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OCEANOGRAPHY AND MARINE BIOLOGY AN ANNUAL REVIEW Volume 20

OCEANOGRAPHY AND MARINE BIOLOGY AN ANNUAL REVIEW Volume 20

HAROLD BARNES, Founder Editor MARGARET BARNES, Editor The Dunstaffnage Marine Research Laboratory Oban, Argyll, Scotland

ABERDEEN UNIVERSITY PRESS

FIRST PUBLISHED IN 1982 This edition published in the Taylor & Francis e-Library, 2005. To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk. This book is copyright under the Berne Convention. All rights reserved. Apart from any fair dealing for the purpose of private study, research, criticism or review, as permitted under the Copyright Act, 1956, no part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, electrical, chemical, mechanical, optical, photocopying, recording or otherwise, without the prior permission of the copyright owner. Enquiries should be addressed to the Publishers. © Aberdeen University Press 1982 British Library Cataloguing in Publication Data Oceanography and marine biology. Vol. 20 1. Oceanography—Periodicals 2. Marine biology—Periodicals 551.46′005 GC1 ISBN 0-203-40060-7 Master e-book ISBN

ISBN 0-203-70884-9 (Adobe eReader Format) ISBN 0-08-028460-4 (Print Edition)

PREFACE Once again it has not been possible to enclose all the manuscripts offered for the present volume. Such enthusiasm to obtain publication in the Annual Reviews is greatly appreciated and must be the best indication of their value to marine scientists. The congenial relations with the contributors and the care of the publishers have again made the editor’s task more rewarding than arduous. The help of everybody is gratefully acknowledged.

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CONTENTS

PREFACE

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Interlinking of Physical and Biological Processes in the Antarctic Ocean D.J.TRANTER The Mediterranean Water Outflow in the Gulf of Cadiz M.R.HOWE The Rôle of Bacteria in the Turnover of Organic Matter in the Sea I.R.JOINT AND R.J.MORRIS Particulate Matter in the Oceans—Sampling Methods, Concentration, Size Distribution, and Particle Dynamics W.R.SIMPSON Biology and Ecology of Marine Oligochaeta, a Review OLAV GIERE AND OLAF PFANNKUCHE The Biology of Sandy-beach Whelks of the Genus Bullia (Nassariidae) A.C.BROWN Recent Studies on the Biology of Intertidal Fishes R.N.GIBSON Aspects of the Bioluminescence of Fishes PETER J.HERRING The Biological Importance of Copper in Oceans and Estuaries A.G.LEWIS AND W.R.CAVE

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AUTHOR INDEX SYSTEMATIC INDEX SUBJECT INDEX

38 74 129

197 351 420 472 534

788 894 929

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INTERLINKING OF PHYSICAL AND BIOLOGICAL PROCESSES IN THE ANTARCTIC OCEAN* D.J.TRANTER Division of Fisheries Research, CSIRO Marine Laboratories, PO Box 21, Cronulla, N.S.W. 2230, Australia

Oceanogr. Mar. Biol. Ann. Rev., 1982, 20, 11–35 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION Present understanding of the remote and inhospitable Antarctic Ocean has developed in an episodic fashion. The pioneering voyages were those of Cook, in the late eighteenth century, and those of Bellingshausen, Wilkes and Ross half a century later. The CHALLENGER, VALDIVIA, BELGICA, GAUSS, SCOTIA and TERRA NOVA played a leading rôle before and after the turn of the century. This was followed by the era of DISCOVERY II when intensive studies were made of Antarctic baleen whales and their environment. After World War II there was the International Geophysical Year (IGY) in which the Soviet vessels OB and VITIAZ were prominent. The momentum of the IGY was continued by the U.S. National Science Foundation vessel ELTANIN. Meanwhile, at laboratories on shore, studies proceeded at a steadier pace, mainly on breeding colonies of birds and mammals and on nearshore communities on the sea floor, in the water column, and embedded in the ice. From this knowledge and understanding has arisen the concept of an Antarctic ecosystem, (Baker, 1954; Currie, 1964; Holdgate, 1967; Knox, 1970; Hedgpeth, 1977). The physical structure of the system has been described by Sverdrup (1933), Deacon (1937), Brodie (1965), Gordon (1971), and Gordon, Taylor & Gingi (1974); plankton productivity by El-Sayed (1968, 1970a,b, 1972), El-Sayed, Mandelli & Sugimura (1964); El-Sayed & Jitts (1973), El-Sayed & Turner (1974), and Holm-Hansen, El-Sayed, Franceshini & Cuhel (1977); the benthos by Knox (1970), Dell (1972), Gruzov (1977), and Knox & Lowry (1977); the ice community by Meguro (1962), Burkholder & Mandelli (1965), Andriashev (1966), Bunt (1966), Buinitsky (1974), and Horner (1974); and the vertebrates by Laws (1977a,b). The references listed here are but an indication of a much wider body of literature. The purpose of the present work is to consider how physical and biological processes interact in the Antarctic Ocean and to discover from the information in hand what forces

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are likely to drive the system. It is directed, in particular, towards “BIOMASS” (Biological Investigations of Marine Antarctic Systems and Stocks), an international collaborative study of the Southern Ocean which is at present in progress. * CSIRO Marine Laboratories Reprint No. 1178.

THE ANTARCTIC ECOSYSTEM The influence of the polar ice cap is all pervasive. Except for those marine birds and mammals that come on shore to breed, there is little life of any kind on the continent itself (Holdgate, 1977); lichens grow on rocky outcrops and, in summer, the snow is sometimes stained by algal growth. Further north, in

Fig. 1.—Distribution of pack-ice about the Antarctic continent in winter (Sept.) and summer (Mar.) 1974: from satellite photography (Budd, 1979).

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Fig. 2.—Seasonal variation in Antarctic sea ice cover (Mackintosh, 1972): in winter and spring, the pack-ice stretches halfway to the Antarctic Convergence.

Fig. 3.—Collection of invertebrates from a bottom trawl in the Scotia Sea, WALTER HERWIG January 1978: the collection includes starfish, brittle-stars, pycnogonids, isopods, gastropods, sponges, octopus, and coral.

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Fig. 4.—The Antarctic krill, Euphausia superba: this species forms the staple diet of many Antarctic animals (photograph by U.Kils).

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milder latitudes, as in the islands of the Scotia Arc, there is a mossy carpet which forms a microvegetation, but this does not sustain any known grazing system (Holdgate, 1967). For the main, the Antarctic ecosystem is the sea. As in the Arctic, the sea is separated from the atmosphere each winter by a layer of pack-ice, which reduces the light available for photosynthesis, shelters the water column from the wind, and cuts off surface aeration. The freezing-melting cycle of the ice controls the dynamics of the water column beneath, through changes in temperature and salinity. A community of algae, photosynthesizing at low light intensities, grows within the pack-ice, providing forage close to shore for a benthic ice community and, out to sea, perhaps, an annual bonanza for pelagic grazers. The seasonal expansion and contraction of the ice cover (Fig. 1) constitute a variable annual pulse to which the components of the system are closely synchronized. The edge of the ice represents a wandering coastline in an open sea (Fig. 2). Most of the species in the benthos are endemic to the Antarctic (Dell, 1972). The fauna has been isolated for a long time, the Scotia Arc providing the main avenue for intermixing. Despite the absence of certain taxa, such as crabs, the benthos is generally abundant and diverse (Tressler, 1964; Knox, 1970; Gruzov, 1977; Richardson, 1977). It is in the shallows where moving ice abrades and scours the bottom of the sea, and in areas bathed by waters from beneath the ice-shelves (Dayton & Oliver, 1977) that the fauna is relatively sparse. Brittle stars, starfish, isopods, amphipods, and sponges are well represented (Fig. 3). The pycogonids, once thought to be rare, are abundant. Most of the fish belong to a single family, the Notothenidae. There are characteristics common to a number of Antarctic benthic species. Most are sessile, sluggish, or slow growing (Holdgate, 1967; Knox, 1970), many are relatively large (Andriashev, 1966), and a high proportion brood their young (Dell, 1972). Their biomass is high but their productivity is low. The hub of the system is Euphausia superba, the Antarctic krill (Fig. 4). This representative of the planktonic Euphausiacea (Crustacea) forms the staple diet of a wide range of fish, cephalopods, birds, and mammals, such as the baleen whales, crabeater seal, fur seal, and Adelie penguin. It is taken even by bottom-living trawl fish (Permitin & Tarverdieva, 1972). Krill is generally considered to feed on phytoplankton but there is evidence (Ikeda, pers. comm.) that it feeds equally well on animal food and, on occasions, is cannibalistic. The system is illustrated in a simple way in Figure 5. One of the components of the ecosystem is Man. Note that both baleen whales and their principal food species (krill) are harvested, but not their principal competitor for forage (crabeater seal).

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LIMITS TO GROWTH The major influence of Man on the Antarctic ecosystem has been through whaling (Gulland, 1976). The stocks of whales are generally much lower now than they were fifty years ago when Antarctic whaling was at its peak (Fig. 6). It is presumed that reduction in whale numbers has lowered the grazing pressure on Antarctic krill. There is evidence that competitors have reaped a benefit from the whales’ demise (Laws, 1977b). Between 1930 and 1960, while the stocks of baleen

Fig. 5.—Simplified Antarctic pelagic food-web, emphasizing the central rôle of Antarctic krill: the crabeater seal is the principal competitor of baleen whales; the Adelie penguin is the principal penguin predator of krill.

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Fig. 6.—Decline in catch of Antarctic baleen whales over the past 50 years (after Gulland, 1976): the fishery moved in sequence from the larger species to the smaller.

whales declined, the number of fur seals at South Georgia increased 1000-fold (Payne, 1977) (Fig. 7). Adelie penguins show a similar trend (Sladen, 1964). After 10 years of commercial whaling in “The Sanctuary”, an area west of the Antarctic peninsula previously closed to whaling, the crabeater seal

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Fig. 7.—Increase in the number of fur seals at South Georgia in relation to the decline of whales during the period 1930–1970 (after Laws, 1977b): the upward trend in fur seal numbers is due in part to reduced exploitation.

began to mature at an earlier age (Fig. 8). Note that neither the Adelie penguin nor the crabeater seal had previously been harvested; the response was indirect. Also, the Sei whale arrived earlier in the summer, penetrated farther south (Gambell, 1968), and its pregnancy rate increased (Gambell, 1973) before the species had been subjected to large scale exploitation. The

Fig. 8.—Progressive lowering of age at first maturity of a population of crabeater seals, following ten years of commercial whaling (after Laws, 1977b).

inference is that this species had begun to benefit from an increasing food supply resulting from the mortality of its larger competitors. In short, the components of the system appear to be food-limited. Baleen whales and crabeater seals are essentially pelagic animals and, as such, have unlimited breeding space—the crabeater seal hauls out on the pack-ice immediately above its food supply and whales migrate to warmer waters. The situation is different with many of their competitors. The breeding sites available to the Adelie penguin, for instance, are those ice-free coasts within easy reach of their staple food supply (krill). Such sites are few and very crowded (Fig. 9). Here, breeding space may well constitute the ultimate limit to population size. Cold per se does not appear to be a major hazard for Antarctic animals. Poikilotherms

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maintain an internal freezing point lower than that of sea water. Invertebrates achieve this by hyperosmotic regulation (Rakusa-Suszczewski & McWhinnie, 1976). Fish which live among the ice have glycoprotein anti-freeze and a high degree of metabolic cold adaptation (Fig. 10) (Feeney, 1974; De Vries & Yuan Lin, 1977, De Vries, 1978). Their problem is heat not cold. They differ from temperate fish which can survive low temperatures in winter and high temperatures in summer. Birds and mammals are well insulated, their problem is how to lose heat in summer on land rather than how to retain heat in winter in the sea. The Antarctic gull solves this problem by using the blood supply to its feet as a heat exchanger (Murrish & Guard, 1977). As far as is known, the main controls on primary production in the sea are nutrients and light. Their availability in the Antarctic for phytoplankton growth is now considered.

NUTRIENTS Figure 11 from the oceanographic atlas of the International Indian Ocean Expedition (Wyrtki, Bennett & Rochford, 1971) illustrates the main features of nutrient distribution in Antarctic waters. Surface nitrate concentration increases with latitude (Fig. 11a), from <0.5 µg-at·1−1 in Australian waters to>10 µg-at.1−1 in the Antarctic. Vertical profiles of inorganic phosphate (Fig. 11b) and silicate (Fig. 11c) between Australia and Antarctica show a similar pattern. The Southern Ocean is relatively rich in nutrients. Concentrations are reduced in summer (Arzhanova, 1974) but there is little evidence that primary production is limited by nutrient availability even though there is little regeneration within a single growing season (Arzhanova, 1976). The source of such high nutrient concentrations is of special interest. When Antarctica and Australia separated and drifted progressively further apart, an uninterrupted passage opened up to the sea (Fig. 12). Within this passage flows the West Wind Drift (Fig. 13), a broad continuous current driven by the westerlies. On its southern flank, adjacent to the Antarctic continent, is a narrow current flowing west, the East Wind Drift, which becomes covered by pack-ice in the winter. Between the West Wind Drift and East Wind Drift is the Antarctic Divergence (Fig. 14) where, under the influence of atmospheric cyclones, warm, deep water, poor in oxygen and

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Fig. 9.—Colony of Adelie penguins, Estneralda Base, Antarctica, January 1978: there are few suitable sites available for penguin rookeries and these are very crowded.

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Fig. 10.—The model proposed by De Vries & Yuan Lin (1977) to explain the antifreeze in the blood of Antarctic notothenid fishes: the glycoprotein component limits freezing by attaching to the surface of the ice crystal.

rich in nutrient, upwells. This is divided by the annual freezing-melting cycle into a lighter fraction (Antarctic Surface Water) and a heavier fraction (Antarctic Bottom Water), both with a drift component to the north. Antarctic Surface Water submerges at the Polar Front beneath the warmer water on its northern flank and continues further northward as Antarctic Intermediate Water. Thus, the water masses of the Southern Ocean consist of a set of more or less concentric circles whose major anomalies are due to bathymetry, coastline topography, and mesoscale features such as eddies. Sverdrup (1933) and Currie (1964) have drawn attention to an “intermediate return current” which could well act as a feedback mechanism concentrating nutrients in Antarctic surface waters (Fig. 15). Low oxygen

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Fig. 11.—Distribution of plant nutrients in the Indian Ocean sector of the Antarctic: surface nitrate (a), phosphate profile to 400 m (b), and

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silicate profile (c) along the transect shown in (a) (dashed line); the units are µg-at.·1−1; concentrations increase progressively towards the Antarctic; after Wyrtki, Bennett & Rochford (1971).

and high phosphate concentration in waters descending sharply north of the Polar Front (Antarctic Intermediate Water) indicate active decomposition and regeneration. Vertical mixing in this area transfers regenerated nutrient across the interface with the southward moving warm deep water and

Fig. 12.—The birth of the Antarctic ecosystem: Australia and Antarctica separate, establishing an open corridor for the Circumpolar Current; after Kenneth (1978).

eventually restores it to the surface by way of upwelling at the Antarctic Divergence. Neshyba (1977) proposed that upwelling from melting icebergs increases nutrient flux through the pycnocline. The experiments of Huppert & Turner (1978) with model icebergs indicate, however, that melt-water

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Fig. 13.—Circumpolar circulation (after Mackintosh, 1973): the prevailing water movement at lower latitudes is towards the east under the influence of the westerly winds; at higher latitudes, close to the Antarctic continent, the movement is toward the west under the influence of the prevailing easterlies; at the interface between the two lies the Antarctic Divergence.

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Fig. 14.—Meridional circulation (after Gordon & Goldberg, 1970): the prevailing circumpolar movement of Antarctic waters has meridional components; cold saline water forms beneath the pack-ice and rolls down the continental slope towards the north while cold, low salinity water spreads northward at the surface; these northward flows are balanced by a southward influx at intermediate levels.

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Fig. 15.—The intermediate return current of Sverdrup (1933): north of the Antarctic Convergence (A.C.), mixing is intense, and there is much mortality and decomposition of Antarctic phytoplankton; part of this Antarctic Intermediate water is entrained by deeper water moving south, thus constituting a nutrient feedback loop.

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Fig. 16.—Mixing and spreading pattern of melt-water from a model iceberg (a small block of fluorescein-impregnated ice in salt-stratified water) (after Huppert & Turner, 1978): melt-water does not rise towards the surface but spreads out laterally at intermediate levels.

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mixes with adjacent sea water and spreads out at intermediate levels instead of rising to the surface (Fig. 16).

LIGHT Of the several somewhat related factors that determine how much light is available for phytoplankton growth, the most obvious is day length. Figure 17 shows the seasonal cycle in day length at various latitudes in the Southern

Fig. 17.—Seasonal variation in day-length at various latitudes: there is little light for photosynthesis in the Antarctic winter.

Ocean. The Antarctic winter appears as a dark wedge interrupting the continuity of photosynthesis. The days get longer and the nights get shorter until, round about the

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September equinox, a phytoplankton bloom develops (Fig. 18). From then until the equinox in March there is as much light available for photosynthesis at high latitudes as there is at low (Fig. 19). In fact, at the surface of the sea, towards the middle of the day, in the Antarctic summer, there is more light available than the plants can safely absorb (Fig. 20). Photosynthesis is highest not at the surface (except on cloudy days) but 5–10 m below (Fig. 21). The amount of light available for photosynthesis is determined not only by how high the sun is in the sky but how high the autotrophs are located in the water column. Because light attenuates with depth, plant production is limited by the depth of surface mixing (Sverdrup, 1953). Evidence is accumulating that phytoplankton production in the Antarctic summer is determined by the stability of the water column (Hart, 1934; Hasle, 1956, 1969; Saijo & Kawashima, 1964; Fogg, 1977).

Fig. 18.—Seasonal phytoplankton cycle as g C·m−2 in surface waters south of the Antarctic Convergence (after Currie, 1964).

It is not the surface light regime that is the proximate driving force but surface turbulence. In turbulent seas, particles are distributed throughout the surface mixed layer (Fig. 22). As the layer deepens, extending farther into the dark, the mean ambient light intensity available to algae living in this habitat progressively declines (Fig. 23). The main determinant of surface turbulence is wind stress. The Southern Ocean is well known for its rough weather and turbulent seas. Figure 24 from the records of the Australian Bureau of Meteorology (Wearn & Baker, 1980) shows the zonal wind stress. The

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Fig. 19.—Diurnal variation in incident solar radiation in the Antarctic compared with regions further north: despite the lower altitude of the sun, the total amount of light on a summer’s day at 64° S is greater than at 31° S; after Holm-Hansen et al. (1977).

Fig. 20.—Diel surface photo-inhibition in the Antarctic summer (Tranter unpubl. data, WALTER HERWIG, Scotia Sea, January 1978, continuous monitoring by Turner design fluorometer): absorption efficiency declines each day at noon (N).

eastern Indian Ocean and west Pacific sectors (south of Australia) have high and variable winds, which induce very deep (≈600 m) mixed layers (Fig. 25). This is the region where “Subantarctic Mode” water is formed (McCartney, 1977), which retains its isothermal character as it drifts northwards at subsurface levels. Such intensive mixing reduces a broad, nutrient-rich, and (superficially) well-lit part of the Southern Ocean to a virtual desert, compared for example, to nutrient-impoverished subtropical gyres. Many workers

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(e.g. Slawyk, 1979) have commented on the poverty of the area

Fig. 21.—Vertical profile of photosynthesis in the Antarctic summer showing subsurface maximum resulting from surface photoinhibition (after Holm-Hansen et al., 1977, Ross Shelf, February 1972).

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Fig. 22.—Vertical profile through the surface mixed layer in the Subantarctic summer, in relation to the light available for photosynthesis (after Slawyk, 1979): paniculate organic matter (including phytoplankton) mixed uniformly to the bottom of the euphotic layer; PN, particulate nitrogen; PC, particulate carbon; T, temperature; NO3, nitrate.

north of the Antarctic Convergence and wondered why production should be lower there than in waters further north where nutrients are limiting. Whereas there is a strong downward component to water movement in the Subantarctic, in some areas of the Antarctic (e.g. the divergence) there is a strong upward component, leading to upwelling. Associated with these upwellings, but on the downstream side (Beklemishev, 1959) there is high

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Fig. 23.—Relative irradiance (473 nm) in various sectors of the Antarctic south of Tasmania and New Zealand (after Matsuike & Sasaki, 1968): the waters south of Australia and New Zealand are extremely clear presumably because there is little phytoplankton.

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Fig. 24.—Zonal wind stress in dynes·cm−2 (after Wearn & Baker, 1980, source:

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Australian Bureau of Meteorology): the areas of greatest (a) and most variable (b) wind stress are stippled—and are most pronounced in the Subantarctic zone, particularly southward of Australia.

primary production. This appears to be a response not to nutrient enrichment as in low latitude areas of upwelling, but to light enhancement associated with the net upward flux. The time lag between upwelling and production is not fully understood (Beklemishev, 1959; Slawyk, 1979). Upwelling occurs not only at the divergence but also in clockwise eddy

Fig. 25.—Vertical section through the Subantarctic (SA) Zone south of Australia showing the deep mixing process which leads to the formation of subsurface “thermostads” (subsurface isothermal layers) (after McCartney, 1977): this process causes extensive phytoplankton mortality; temperatures in °C).

systems such as those in the Weddell Sea (Foster, Carmack & Neshyba, 1976) to the northeast of the Ross Sea, and in the vicinity of the Kerguelen-Gaussberg Ridge (Gordon, 1971). Walsh (1969), El-Sayed & Turner (1974) and Holm-Hansen et al. (1977) question the “proverbial richness” of Antarctic open-ocean waters. It is usually nearshore that high productivity is observed (Bienati, Comes & Spiedo, 1974).

THE PACK-ICE SYSTEM The layer of pack-ice which forms on the surface of the open ocean around the Antarctic continent adds a new dimension to the Antarctic ecosystem. The pack-ice provides a relatively stable substratum for micro-algal growth which, in turn, radically alters the physical properties of the substratum. Thus the pack-ice is a system within a system, or a seasonal phase of an essentially two-phase ecosystem: the summer phase, predominantly pelagic, the winter phase with strong benthic or “continental” overtones. Although the growth of micro-algae in polar pack-ice has been known for over 100

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years, the full significance of its rôle in Antarctic productivity has only recently begun to emerge (Meguro, 1962; Burkholder & Mandelli, 1965; Horner, 1974). As icebreakers move through the pack-ice in the summer, this algal layer is seen as a brown discolouration almost everywhere. The chlorophyll concentration of the layer is of the order of 0.1 g·m−2 (≈3.7 g C·m−2, assuming a carbon: chlorophyll ratio of 37:1, Bunt & Lee, 1972), and the area covered by the annual expansion and contraction of the ice is ≈ 15+106 km2. As this is the habitat of Euphausia superba, the rôle of “ice” algae in the krill ecosystem may well be significant. Daily phytoplankton production in the ice-free water column of the open ocean during the growing season (≈150 days) is of the order of 0·1 g C·m−2·day−1 (Burkholder & Mandelli, 1965; Holm-Hansen et al. 1977) i.e. 15 g C·m−2·yr or four times the standing crop of algae in the ice. Because this is released into the water column in a concentrated form, as the pack-ice breaks up and retreats toward the Antarctic continent, it represents forage that is available for grazing at low energy cost. Whether this is done, and how, and by what grazing populations is not yet known. There is, however, historic evidence that this is where the pre-war stocks of baleen whales used to aggregate. The rôle of “ice” algae in the krill ecosystem merits urgent investigation. The most plausible model of the development of algal growth in the pack-ice is that of Meguro (1962) (Fig. 26). In autumn, the air is much colder than the sea and pancake ice forms on the sea surface. This collects and retains the

Fig. 26.—Model for algal growth in Antarctic sea ice (after Meguro, 1962): Pancake ice forms on the surface of the sea in autumn (1) and collects the winter snow (2) which submerges the ice allowing sea water to penetrate at the ice-snow interface (3); within this matrix an algal mat (A) develops, when there is enough light for photosynthesis (4); in summer, the algal mat absorbs so much heat (5) that the “pack-ice” melts and breaks apart liberating the entire algal crop into the water column (6).

autumn snowfall which in the course of time becomes heavy enough to submerge the ice. Thus, high-nutrient sea water is allowed to percolate along the interface between ice and snow forming a porous matrix in which there is enough light and nutrients for photosynthesis. In spring, growth resumes again and absorption of sunlight by the brown pigmented layer so weakens the pack-ice at the algal interface that the top half breaks away releasing the algal crop into the water column. There are other ways in which algal growth can develop in the ice, particularly near to shore, but the areas involved are so much less than the area of the pack-ice that this

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contribution to overall Antarctic production cannot be as great. In addition, the growth close to shore frequently takes place on the lower surface of the ice where there is less light available for photosynthesis. This “epontic community” first described in detail by Bunt (1964, 1966) is adapted for photosynthesis at low light intensities. There may be a “spring epontic community”, which develops on the lower surface of the nearshore ice in spring, and an “autumn epontic community”, which becomes incapsulated in the ice (Hoshiai, 1974). Buinitsky (1974) found that the concentration of algae in the inner layer of the ice is often greater than that in the lower surface layer of newly formed ice, presumably because higher up more light is available. Volume for volume, cell concentrations in the ice throughout the entire period of ice cover (April–January) were 1–2 orders of magnitude higher than in the water column below. Andriashev (1966) and Gruzov (1977) describe graphically how these nearshore layers form. The ice crystals rising from the bottom are colonized by diatoms as they rise and become incorporated in the autumn ice as an autumn algal layer. With the summer thaw, this rich organic mass is liberated into the water and much of it falls out on the shallow bottom, providing forage for detritivores and grazers. Now more light is available for the growth of algae on the sea floor. These survive until the following autumn when ice forms once again and scours the bottom clean. In the Weddell Sea, layers of algae are found within the ice, associated with brine maxima (Ackley, Buck & Taguchi, 1979). These are thought to form in response to seasonal changes in the porosity (and, therefore, the buoyancy) of the ice, where nutrients are available for photosynthesis. Little is known about the nutrient supplies available to “ice” algae. Meguro’s model (Fig. 26) implies that nutrients reach the “snow community” with sea water percolating laterally through the algal matrix. It is likely that, in the long summer days, growth is limited by nutrients rather than by light. By contrast, as salt is excluded from the surface ice, the water column below continues to mix by convective overturn (Gordon, Taylor & Gingi, 1974) and, simultaneously is shaded by algal growth in the ice above. Here, light is more likely to be limiting than nutrients. Burkholder & Mandelli (1965) found production in this shaded habitat to be only one thirtieth that of the algae in the ice. The effect of “ice” algae on the physical properties of the ice in which they are embedded is described by Buinitsky (1974). Shipboard tests of ice strength under compression and under horizontal stress showed that ice with algae in it was only half as strong as ice taken from the same block without any algae, the degree of weakness being a function of the algal concentration. Meguro (1962) has drawn attention to the further weakening which takes place in summer as the pigmented algal layers absorb heat faster than the layers between which they are sandwiched. The effect of algae on the chemistry of the ice is described by Dradovskiy (1974). Nitrite concentration is high in the surface (snow-ice) algal layer of the pack-ice, suggesting that there is growth and remineralization of organic matter in situ. This is consistent with the Arctic studies of Grainger (1974) who showed that nutrient levels in the ice are progressively reduced, presumably as a result of algal growth. The “ice” algae observed by Buinitsky (1974) were alive and, in some cases, actively reproducing. Horner (1974) has reviewed the recent work in this interesting and important field.

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THE HABITAT Within the circumpolar circulation there has developed a unique Antarctic fauna (Hedgpeth, 1969, 1977). Its northern boundary is the Subtropical Convergence, corresponding approximately to 40° S latitude. This varies considerably from season to season, and from one meridian to another. The Antarctic Convergence or Polar Front is a meandering interface embedded in the West Wind Drift where there is active interchange of heat and salt, in part by way of eddies (Gordon et al., 1974; Gordon, Georgi & Taylor, 1977; Joyce & Patterson, 1977; Sievers & Emery, 1978). The productivity here is high, due perhaps to the fact that although the water column is structurally complex it is also relatively stable. Despite a strong meridional component to the predominantly circumpolar circulation, this Polar Front is a major biogeographic barrier. Mackintosh (1937) has shown that some Antarctic macrozooplankton migrate seasonally between surface waters which have a northerly drift component (summer) and deeper waters which have a southerly component (winter). According to Vladimirskaia (1975), 60–70% of the winter zooplankton biomass is concentrated below 200 m. Such a seasonal migration pattern conserves plankton populations south of the Antarctic Convergence and adds to its effectiveness as a biogeographic barrier. The region south of the Antarctic Convergence consists of the “low Antarctic”, within the West Wind Drift, and the “high Antarctic”, within the East Wind Drift. It is in the East Wind Drift and its associated gyres and eddies that Euphausia superba breeds (Marr, 1962), i.e. in the area covered by pack-ice in the winter. The most important of these mesoscale features is the cyclonic Weddell Gyre, formed in the south Atlantic sector by the interaction of the East Wind Drift with the northward sweep of the Antarctic Peninsula and the islands of the Scotia Arc (Beklemishev, 1959; Carmack & Foster, 1974; Deacon, 1976; Foster et al., 1976). According to Marr (1962), krill eggs sink, and hatch at depth, the larvae swimming upward through the warm deep current at the nauplius stage, and reaching Antarctic Surface Water at the first calyptopis stage. Thus they would pass first through waters with a northerly drift component (Antarctic bottom water) and continue their development in waters with a southerly drift component (warm deep water). Spawning has usually been observed to take place at fronts (Makarov, 1972), areas of upwelling (Mackintosh, 1972), for example the Antarctic Divergence, and on the continental shelf (Hempel, Hempel & Baker, 1979). Voronina (1974) has drawn attention to the consequence of eggs falling beyond the point of no return, i.e. beyond the depth where larval yolk reserves can sustain the nauplius throughout its long climb back towards the surface where algal food is available. The expatriation range of the species extends northward into the West Wind Drift, e.g. to the Kerguelen-Gaussberg Ridge and, in the eastern Pacific, into Chilean fjords. In winter, the entire breeding range of the species, and the greater part of its expatriation range, is covered by pack-ice. Hence, the populations are exposed for only a few months in the summer. The period of exposure is shorter in the Pacific than in the Atlantic or Indian sectors, and particularly short to the south of Australia. Fordyce (1977) presents evidence which indicates that evolution of the baleen whales

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took place in mid-Oligocene times when the Circumpolar Current was born. Today they follow the annual retreat of the pack-ice towards the Antarctic coast, and feed on krill newly exposed to predation (Mackintosh, 1973). Pre-World War II stocks of whales moved on, in March, into the East Wind Drift; post-war stocks remain in the West Wind Drift, where, presumably, the forage is now adequate for their needs. As the ice breaks up, the algal crop is liberated into the water column where it is available to the krill and to the summer crop of phytoplankton. It is not yet known whether Euphausia superba make use of it; this constitutes one of the most important questions in the field of Antarctic productivity which need to be addressed. Antarctic krill has generally been regarded as an open ocean species, but it is as much a creature of the ice. The Antarctic ecosystem has been thought to be principally pelagic. Its benthic character (Bunt, 1966) merits wider recognition. Animals over 20 mm in length frequently occur in swarms, detectable by echo sounder and sometimes visible from the surface. According to Marr (1962) these swarms range in extent from several hundred metres to several kilometres. Although the shape of the swarm continuously changes, in Marr’s opinion it behaves as a unit, analogous to a shoal of fish. Very little is known about the physical and biological factors which control swarms of krill. As these may be the ultimate grazing units for baleen whales, seals, and even birds, their ecology is of particular importance. Of the total zooplankton biomass in the water column in winter, 60–70% is concentrated below 200 m (Vladimirskaia, 1975). If Euphausia superba share this habit, it is likely that upwelled waters would bring them to the surface in the summer. According to Elizarov (1971), however, krill swarms coincide not with the actual sites of upwelling but with adjacent areas where waters sink. There is a burst of biological activity followed by a population explosion as krill increase in size and reproduce (Latogurskii, Naumov & Pervushin, 1975). Is the observed sparsity of phytoplankton at upwelling sites due to grazing, then, or to time lags in primary production or both?

CONCLUSIONS The ocean south of the Antarctic Convergence contains a circumpolar community of plants and animals characterized by Euphausia superba, the Antarctic krill. Birds, seals, whales, and other animals use krill for food, and there is evidence (for example increased reproduction rate) that present populations now enjoy a more generous food supply than when the stocks of baleen whales were large. This indicates that the biological resources of the Antarctic must be understood and managed as a multi-species system. Primary production in the Antarctic is limited primarily by light. There is little light for photosynthesis in winter because the days are short and, south of latitude 60°, the surface of the sea is covered by ice and snow. In summer, where the sea is ice-free, the wind stress is frequently so great that the phytoplankton is mixed well beyond the compensation depth where gains by photosynthesis are lost in respiration. There is, however, an extensive growth of “ice” algae living near the waterline of the sea ice, at the interface between ice and snow. Release of this organic matter into the water column when the ice breaks up in summer constitutes a major event in the annual production

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cycle of Antarctic waters. It is here that the large pre-World War II stocks of baleen whales used to feed. Nutrient concentrations in Antarctic waters are so high that they are unlikely to limit primary production except, perhaps, in the algal matrix of sea ice and in the surface layer of melt-water. The high phytoplankton production sometimes observed in areas of divergence appears less likely to be due to nutrient enrichment than to light-associated factors related to the wind regime. The faunas on either side of the Antarctic Convergence are distinctly different, their separate identities persisting despite the prevailing northerly drift of Antarctic water at intermediate levels. There is evidence that this is due to seasonal vertical migration between surface waters with a northerly drift component, and relatively warm deep waters with a southerly component. Sverdrup’s “intermediate return current” could constitute a feedback mechanism enhancing nutrient concentrations south of the Antarctic Convergence. It is important to abandon the concept that the Southern Ocean is universally productive and to concentrate on processes taking place in the pack-ice and at the Antarctic Divergence and Convergence. Attention should be focused on the light regime, both in relation to the stability of the ater column and to the growth of “ice” algae, on the ecological balance between baleen whales and their smaller competitors such as seals, and on the effect of the emerging fishery for krill upon the recovery of the remaining stocks of baleen whales.

DISCUSSION The main conclusion arising from this review is that the Antarctic ecosystem is foodlimited and the driving force is light. Light, in turn, has two main determinants, daylength and wind stress, the latter influencing light through turbulence and mixing. It would be rewarding to explore the effect of light in the Antarctic ecosystem, not only as a driving force, governing primary production, but also as a control mechanism synchronizing life history cycles with the seasonal production cycle. There is evidence, for instance (Griffiths, Seamark & Bryden, 1979; Bryden, pers. comm.), that the rate of change of day-length in the spring is the factor that stimulates gonad development in the elephant seal, the effect being mediated via the pineal gland which is well developed in polar animals (Cuello, 1970; Piezzi, 1973). Fine tuning of this kind may well be crucial for those Antarctic animals which “can feed where they wish but must breed where they can” (Murphy, 1964). Not only do they have to survive from one summer to the next and store up reserves for breeding but, in addition, they must time their breeding cycle so that their young can gather enough food in their first summer to survive the following winter. The methods used by Antarctic animals to survive the winter are of special interest. Whales commute annually between the Antarctic and the tropics. Some species store food reserves, hibernate or feed on the survivors (Gruzov, 1977). Their reproductive pattern reflects the seasonal variation of their habitat. Many species brood their young (Dell, 1972). Antarctic animals are generally larger than their low latitude counterparts and longer lived (Andriashev, 1966; De Broyer, 1977); they have the chance to breed a

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second time, a property of great survival value in a highly seasonal habitat. They escape in time, as it were, as other (migratory) species escape in space. For the main, the Antarctic winter, however, constitutes a blank in our understanding of the Antarctic ecosystem, particularly the world beneath the ice, the winter habitat of krill. There is urgent need to fill this gap. The problems are logistic. Observations must be made through the pack-ice well out from land, preferably near the ice edge. A suitable platform would be a vessel built like the FRAM overwintering in the ice. Fifty years ago, pelagic whaling got underway in the Antarctic on a massive scale and scientific research was done to monitor its effects. Now we are at the threshold of another exploitative era, the target species including not only whales but also their staple food supply, Antarctic krill. The science of managing multispecies systems is in its infancy. It would be difficult to nominate an optimum harvesting plan for even a simple system involving only whales, crabeater seals, and krill (Fig. 5). Table I suggests that of the various options involving these three species, the least desirable is the one that is most likely to eventuate. This illustrates the need for gathering more information about how the Antarctic system functions, and how its major biological constituents interact with each other and with their common physical environment. If the North Pacific salmon now established in Chilean waters (Joyner, 1980) should start to feed on krill, a major new variable could enter the equation.

TABLE I Harvesting options for Antarctic marine resources in order of their likelihood of maximizing and sustaining the stocks. 1.

Harvest crabeater seals

2.

No harvest of seals, krill or (baleen) whales

3.

Harvest seals+krill

4.

Harvest krill

5.

Harvest whales+seals

6.

Harvest whales

7.

Harvest whales+krill+seals

8.

Harvest whales+krill

The current international collaborative research programme known as “BIOMASS” is directed at such questions. This programme is being coordinated by a working group of specialists under the auspices of SCOR, SCAR and ACMRR. The first major field work (“FIBEX”=First International Biomass Experiment) took place in the Austral summer of 1980/81. The management Commission foreshadowed in the Convention for the Conservation of Antarctic Living Resources will depend heavily on the information and understanding generated by BIOMASS.

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THE MEDITERRANEAN WATER OUTFLOW IN THE GULF OF CADIZ M.R.HOWE Oceanography Department, University of Liverpool, Liverpool, England

Oceanogr. Mar. Biol. Ann. Rev., 1982, 20, 37–64 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION If one were to refer to the Meteor Atlas (Defant, 1936), which illustrates so impressively the hydrographic observations that were made by the German Atlantic Expedition of 1925–1927, the tongue-like spread of the Mediterranean water into the North Atlantic would seem to emanate from an apparent “epicentre” near Cape St Vincent (37°N:9°W) rather than from the Strait of Gibraltar itself (36° N:6° W). The maximum values of salinity and temperature that are quoted for this “transposed source” occur at a depth of 1000m and are somewhat in excess of 36.4‰ and 12 °C. This representation of the Mediterranean water intrusion into the Atlantic therefore suggests that, in the initial stages of the outflow there is a preference for the undercurrent to follow a fairly direct and restricted route towards Cape St Vincent with little lateral dispersion, and it is only beyond this point that the more uniform tongue-like divergence continues into the open ocean. This review will concentrate on recent research, particularly during the period 1970– 1980, which has attempted to determine the characteristics of the Mediterranean undercurrent in these initial stages of the outflow, and our attention will, therefore, be confined to the results of work that has been done in a region extending from the Strait of Gibraltar, through the Gulf of Cadiz, to the south of Cape St Vincent (Fig. 1). Previously there had been several expeditions by ships of different nationalities to this area, but they made comparatively few measurements and these were only at certain isolated locations. In 1958 the British R.R.S. DISCOVERY II, and more significantly in 1967 the French research vessel JEAN CHARCOT, undertook what must be considered as the first large scale surveys that were designed to study the overall flow pattern of the Mediterranean undercurrent. A preliminary report of the French cruise by Lacombe, Madelain & Gascard (1968) summarizes the principal aims of that ex-pedition, which were to make hydrographic sections and current measurements at certain strategic positions that would hopefully intercept the main flow paths of the Mediterranean water in its passage through the Gulf of Cadiz. The results were interpreted by Madelain (1970)

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and in a very comprehensive discussion he was able to include data from both the above cruises as well as from several shorter surveys that were made by other ships during the period 1957–1967. Madelain’s account begins with a description of the undercurrent as it emerges from the Strait at about 6°20′ W where he recorded maximum salinities and temperatures of the order of 38.28‰ and 13.25 °C. His main conclusion was that the outflow, after turning sharply to the right due to the effect of the Coriolis force, would thereafter be considerably influenced by the unusual nature of the sea floor topography. This was known to be quite rugged due to the presence of several prominent submarine channels, shelf canyons, and sea mounts. As a result Madelain proposed that the undercurrent would be divided into two main streams. The first remains in contact with the Spanish continental shelf, whereas the other is subsequently formed by the rapid offshore flow of much of the original water mass down two of the shelf canyons. These branches of the flow would then merge at about 8° W (Fig. 1) to produce the main westward moving current that is

Fig. 1.—The Mediterranean outflow in the Gulf of Cadiz (after Madelain, 1970): the principal hydrographic observations and current meter data were recorded along Sections I to VII during the cruise of JEAN CHARCOT in 1967.

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eventually responsible for the general spread of this water mass into the open Atlantic. There was, of course, no opportunity to investigate any variability in the outflow regime from such a set of widespread and non-synoptic data, nor indeed was it possible to appreciate the degree of inhomogeneity in the Mediterranean water structure. Madelain did note, however, with much interest, a distinct subdivision of the outflow at about 8°30′ W that took the form of separate maxima in the temperature and salinity profiles. For example, he recorded values of 3640‰ and 12.33 °C, which constituted a maximum at the relatively shallow depth of 858 m. These can be compared with a salinity of similar magnitude, 36.60‰, but lower temperature, 11.82°C, which were typical of the values that he observed in the main Mediterranean core at about 1300 m. Similarly Swallow (1969), in an analysis of the Discovery hydrographic sections, was able to intermittently identify Mediterranean water of variable intensity between depths of 675 m and 775 m, where again the salinity values were sometimes found to be equivalent to those normally observed in the deeper layers. These apparent anomalies in the structure of the undercurrent were soon to attract considerable attention. There is, however, no doubt that Madelain’s overall assessment of the outflow regime has been used as the basis and inspiration for the more detailed investigations that have taken place since 1970. These have included similar large scale surveys as well as longer hydrographic time series and current measurements in more advantageous positions. In some cases the interpretation of the results has been accompanied by certain modifications to the flow pattern while, at the same time a great deal of new information has become available concerning the structure and dynamics of the undercurrent. This in turn has initiated an interesting discussion regarding the causes for the high degree of inhomogeneity in the outflow and thereby focused attention on the various mixing processes that may occur within or near the Strait of Gibraltar itself.

OBSERVATIONS IN THE MEDITERRANEAN OUTFLOW 1970–1980 It was the availability of continuously recording salinity-temperature-depth systems that first revealed the true complexity of the Mediterranean outflow structure. This was originally appreciated when measurements were being made, not within the Gulf of Cadiz itself, but in an area about 40 miles to the west of Cape St Vincent. In 1967 German oceanographers, using a bathysonde type of instrument (temperature-electrical conductivity-pressure), quite frequently observed a division of the Mediterranean water into two cores which could be identified as separate maxima in the vertical profiles of temperature and salinity. In addition, the instrumentation was able to resolve a considerable amount of fine structure of varying vertical scales through a large part of the water column. The first reports of these new results by Zenk (1970) and Gieskes, Meincke & Wenk (1970) described some of the physical characteristics of the upper and lower cores which were located at depths of 750 m and 1170 m, respectively. During a ten-day time series at an anchor station Zenk noted a considerable variation in the heat and salt content within the layers occupied by the Mediterranean water, but basically it was possible to distinguish in his mean temperature-salinity profile values of 11.8 °C and

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36.18‰ for the upper maximum and 11.3 °C and 36.38‰ for the lower maximum. At the same time Howe & Tait (1972) were also recognizing the existence of these separate temperature-salinity maxima in this region. They made an analysis of the fine structure that was associated with both the Mediterranean cores and which was represented in the profiles by a large range of inversions of different vertical scales. The stability of such anomalous features in the ocean was of particular interest. In a comparison of the physical characteristics of the inversions in the two maxima it was concluded that the stability generally increased for the smaller scales, but when considering equivalent scales it was consistently greater for the inversions in the upper core. This tendency was similarly reflected in the mean stability profile for the region where the values at a depth of 600 m were almost twice those at depths of about 1200 m. It was, therefore, evident from the above studies that two separate cores, with quite discernible differences in their physical properties, could be easily distinguished in the Mediterranean undercurrent at a distance of 200 miles from the Strait of Gibraltar, and any explanation of this phenomenon would necessarily entail more precise measurements nearer the source. In the following years various researchers were, therefore, motivated by the thought that observations within the Gulf of Cadiz itself would be particularly productive, especially because the modern continuous recording salinity-temperaturedepth instrumentation was now readily available. As a result several cruises have been made to the area, each concentrating on different aspects of the outflow. In 1971 the F.S.METEOR conducted a large scale current measurement programme while at the same time the R.R.S. DISCOVERY was carrying out some variability studies. In 1973 the R.R.S. SHACKLETON and in 1976 the R.R.S. CHALLENGER both undertook large scale hydrographic surveys, and it can be claimed that all these expeditions made major contributions to our present knowledge of the intrusion of the Mediterranean water into the Gulf of Cadiz. It is probably true to state, however, that these efforts relied a great deal in their planning stages on Madelain’s (1970) observations and interpretation. This review will now consider the results from these cruises, as well as other data and theoretical work, not necessarily in chronological order but rather when appropriate to the discussion of such topics as the overall flow pattern, mixing, variability, and velocities in the undercurrent.

HYDROGRAPHIC SURVEYS R.R.S.SHACKLETON CRUISE 1973 This was a significant cruise in two respects. First, it provided an opportunity to confirm some of the principal features in Madelain’s (1970) outflow pattern while using a modern salinity-temperature-depth system. Secondly, by collecting water samples at certain depths and applying the latest analytical techniques it was hoped to establish for the first time the general nutrient salt distribution throughout the region. The results of the combined analysis of both the physical and chemical surveys proved to be particularly interesting. Four sections of hydrographic stations were worked, with three positioned along the submarine canyons which are a prominent feature of the bottom topography

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between the Strait of Gibraltar and Cape St Vincent. These were selected primarily because Madelain had decided that they were exerting a considerable influence on the flow paths of the Mediterranean water. In fact the stations in these sections coincided very closely with those of the Jean Charcot cruise which are shown in Sections IV, V, and VII of Figure 1. In general, the large scale characteristics of the outflow, as recorded by both the Jean Charcot survey in 1967 and the Shackleton survey in 1973, appear to be virtually constant. Ambar, Howe & Abdullah (1976) have given a full account of the chemical and physical data from the Shackleton cruise and Figure 2 from that paper, together with Figure 3 (Madelain, 1970) have been reproduced here to demonstrate the apparent long term consistency in the temperature and salinity values along Section IV (Fig. 1). Furthermore, by reference to the original publications, the reader can

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Fig. 2.—Distribution of salinity and temperature along a section coincident with IV in Fig. 1 (after Ambar et al., 1976).

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Fig. 3.—Vertical sections of salinity and temperature along Section IV (see Fig. 1) (after Madelain, 1970).

make a similar comparison of the respective sets of observations that were made along Section VII (Fig. 1) during these cruises. The sudden turn to the right as the outflow leaves the Strait and its subsequent movement along the Spanish continental slope with little offshore spreading, is easily recognizable. It is perhaps not surprising that with such a substantial and continuous outflow of this kind the temperature-salinity values near the Strait should remain so constant. The less predictable fact that the offshore spreading appears to remain so restricted must, however, also contribute greatly to the consistency of the flow characteristics. The first difference in interpretation arises when Ambar et al. (1976) consider the large scale structure of the undercurrent. They refer to the layer of Mediterranean water that settles out at depths above 900 m as the “upper core” and then proceed to treat it as a separate water mass as it moves westward. This was justified to a large extent by the extraordinary variation in the nutrient values that was revealed in the chemical analysis of the water samples. Consequently, Howe, Abdullah & Deetae (1974) have suggested that the source of this shallower Mediterranean core might be more appropriately associated with depths within the Strait of Gibraltar that are significantly different to those from which the main outflow usually descends. This conclusion was

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based on the combined study of the physical and chemical data which essentially showed that whenever this upper core was sampled in the Gulf of Cadiz, there were anomalously low concentrations of the nutrient salts nitrate, silicate, and phosphate. So much so that when the chemical samples coincided with the actual upper temperature-salinity maximum at depths between 600 and 800 m, the nutrient values were not only less than those of the usual source of the outflow (that is, between depths of 150–300 m in the Strait) but they were also less than the values that can be normally observed in the layers of the Atlantic through which any mixing would occur. In other words, there seemed to be no way of accounting for these abnormally low concentrations in the upper maximum by reference to either the source water or the ultimate mixing environment. Howe et al. (1974), therefore, decided that an explanation of this chemical anomaly might provide the means of identifying the source of the upper Mediterranean core. The relevant information is conveniently presented in Figure 4 where the salinity-nutrient relationships have been plotted for the various water masses that are likely to be involved in the intermixing. These include a profile for the source waters, which were sampled just within the Mediterranean Sea itself; the computed regression lines for the samples that were collected at the standard depths of 600 and 800 m in the Atlantic where the upper core is to be found; and the average salinity nutrient values for the normally undisturbed layers in the Atlantic through which the upper core would have to mix in its initial stages of intrusion. The authors then focus attention on the high salinity (low nutrient) values labelled X and Y (Fig. 4) which represent the properties of the upper Mediterranean core as it was observed in its most original or pure state. They argue that when the possible mixing lines between the various layers of the water column in the Mediterranean Sea and the Atlantic are considered (see caption to Fig. 4) the points X and Y, with their extremely low nutrient concentrations, cannot be produced by outflow water emanating from the usual depths of 150–300 m, nor from a completely mixed water column between the surface and 300 m. The only suitable source water, with concentrations most appropriate to the likely mixing lines (broken in Fig. 4), is that from depths above 150 m, and more precisely from a depth of 110 m. This water would have to pass through the Strait with virtually no vertical mixing in order to account for the nutrient concentrations that were observed in the upper core. This interpretation, however, is not entirely compatible with a theory that Siedler (1968) proposed, and which Zenk (1970) applied to his earlier observations of the double temperature-salinity maxima off Cape St Vincent. Siedler postulates that such a double maxima can be readily generated by the tidal mixing processes within the Strait of Gibraltar. In his model he uses published data to represent approximately the current and salinity structure, and by superimposing the effects of a parabolic velocity profile, a surface tide and an internal tidal boundary wave, he produces a frequency distribution of water types with two preferred salinity values. In Zenk’s application of this theory it is assumed that these two Mediterranean water types will emerge from the Strait and, having mixed with North Atlantic Central Water at an initial depth of 500 m in similar ratios, will then form the two distinct cores in the outflow which he had observed to the west of Cape St Vincent. Unfortunately it is the degree of tidal mixing within the Strait, which is so implicit in Siedler’s model, that is inconsistent with the chemical interpretation suggested by Howe et al. (1974). In spite of this, the information being accumulated from these studies strongly

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indicated that a Mediterranean upper core can be formed, and thereafter it would survive as a separate water mass at distances of at least

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Fig. 4.—The regression relationship (solid lines) for salinity-nutrients in the Atlantic at 600 and 800 m and the salinity-nutrient profile (heavy line) at a station within the Mediterranean Sea: average salinitynutrient content of the Atlantic (A) and Mediterranean (M) water columns between the following depths: A1:0–800 m; A2:0–600m; M1:0–150m; M2:50–150 m; M3:0–300 m; M4:150–300 m; for the relevance of the broken lines and points X and Y see p. 43; (after Howe et al., 1974).

200 miles from its probable source. Accordingly Ambar, Howe & Abdullah (1976) proceed to examine their Shackleton hydrographic data with the principal aim of identifying, wherever possible, this kind of stratification in the outflow. They start by again referring to the canyon section in Figure 2, and discuss the ultimate fate of the ‘anomalous’ mass of Mediterranean water near the bottom of Station 92, where a salinity of 36.88‰, a potential temperature of 12.95 °C and a potential density of 27.88 were recorded at a depth of 870 m. More typically such densities are observed in the deeper water further offshore where, for example, at 1200 m in Station 98 the salinity was 36.52‰ and the potential temperature was 11.63 °C. They then pose the question: does part of this water flow down the canyon and thereby initiate the main deep core of the outflow or does it remain isolated at the shallower depth and make no contribution to the colder and less saline water mass that is to be observed between Stations 97 and 98? Ambar et al. (1976) point out that the salinity-nutrient relationship at a depth of 800 m in Station 92 is closely correlated with that of the upper core water mass in the other sections further to the west (Fig. 5). They, therefore, propose that because of the prevailing ambient conditions this water mass will lose salt fairly rapidly due to the large horizontal salinity gradients, while maintaining its relatively high temperature, and ultimately it will achieve a potential density of about 27.49. This will then account for its presence at a depth of 793 m in Station 71 (Fig. 5) with a salinity of 36.52‰ and a potential temperature of 13.48 °C. In contrast, the lower Mediterranean core at a depth of 1270 m in Station 73 of this same section has a salinity of 36.59‰, a potential temperature of 12.09 °C, and a potential density of 27.82. Finally, and to emphasize the amount of stratification that appeared in these Shackleton sections, it was possible to discern yet another temperature-salinity maximum in the profiles. This occurred at the relatively shallow depth of 400–500 m but only in the inshore stations at the head of the shelf canyons (Figs 2 and 5). If these profiles had been examined in isolation and without reference to the properties of the surrounding waters, it would have been tempting to speculate that this maximum was due to a further subdivision of the shallow core of the Mediterranean undercurrent. When this feature was, however, considered in relation to the general temperature and salinity distribution, it seemed more likely that it had been caused by the sinking of the surface layers which, according to Ambar et al. (1976), had been induced by the winter cooling of the shelf waters and facilitated by the presence of the canyons. If one were to summarize the evidence that has been presented up to now, it would be reasonable to assume, first, that there is indeed a strong preference for the outflow to maintain contact with the Spanish continental slope in its progress towards Cape St

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Vincent. Secondly, the long term constancy in the downstream values of the temperature and salinity denotes very little variation or patchiness in the overall flow regime. The most significant new development in the study of the undercurrent has been to establish the existence of a separate Mediterranean water upper core, which is characterized by a consistently higher temperature than that of the main lower core, and abnormally low nutrient concentrations. It can be initially identified fairly near to the source in Section IV (Fig. 1) and then clearly distinguished in Section VII. This presumably ensures its permanent presence in the area and, therefore, readily accounts for the upper temperaturesalinity maximum in the profiles that were originally observed in the open Atlantic to the west of Cape St Vincent (Zenk, 1970; Howe & Tait, 1972).

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Fig. 5.—Distribution of salinity and temperature along a section coincident

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with VII in Fig. 1 (after Ambar et al., 1976).

R.R.S. CHALLENGER CRUISE 1976 It was acknowledged that the Shackleton 1973 cruise was somewhat exploratory and so the same scientific personnel considered it important that the Challenger 1976 survey should at least confirm some of the previous concepts concerning the structure and the general flow pattern of the Mediterranean undercurrent. The two cores were now believed to be a permanent feature of the outflow and so the zone in which the separation occurs needed to be identified, as well as their preferred routes through the Gulf of Cadiz. In addition, their rate of mixing and any significant changes in the thermohaline characteristics would be of interest, together with estimates of the velocities along the flow paths. With these aims a very comprehensive grid of stations was worked from Cape St Vincent to the entrance of the Strait of Gibraltar, and again this survey included a series of sections along the prominent canyons. The results have been published by Ambar & Howe (1979a, b) and later they will be discussed in some detail under more appropriate headings. It was, however, immediately obvious that there was no significant intermittency in the outflow pattern and that the maximum temperature-salinity values associated with both cores were easily discernible throughout the entire region. For example, in a north-south section off Cape St Vincent (Fig. 6) which was, of course, at the western extremity of our area of interest, the large scale stratification was plainly visible. Here Ambar & Howe (1979a) recorded maximum values of 36.57‰ salinity and 12.18°C in the lower core at a depth of 1315 m with a potential density of 27.83, whereas in the upper core the maxima at 755 m were 36.43‰ and 13.07°C with a potential density of 27.52. In addition, they also noted the presence of an even shallower temperature-salinity maximum in the profiles of the inshore station at the edge of the continental slope, where the values at 580 m were 36.03‰ and 12.90°C with a potential density of 27.24. This water mass is something of an enigma because although in the previous Shackleton survey, which was made during April 1973, this feature had been attributed by Ambar, Howe & Abdullah (1976) to winter cooling and the subsequent sinking of water from the shelf, the appearance in August of a similar subsurface layer makes their explanation more dubious. Instead, Ambar & Howe (1979a) raise the possibility that this might be a further shallow subdivision of the undercurrent. The idea gained support when they were able to identify the probable point in the outflow, at about 7°W and near Section III (Fig. 1), where the separation of the two main cores occurs. Figure 7b shows this section and it can be conveniently compared with a section through the undivided outflow (Fig. 7a) as it leaves the Strait. In Figure 7b Ambar & Howe assign maximum values of 37.42‰, 13.16°C, and a potential density of 28.28 to the main lower core at 756m; 37.07‰, 13.72°C, and a potential density of 27.88 for the upper core at 650 m; and values of 36.65‰, 13.61°C, with a potential density of 27.57 at 510 m in the inshore station where the shallower subdivision of the upper core might be occurring. If so, this shallow water mass would then have to follow closely the 500 m isobath in order to appear as the anomalous maximum at 580 m in the section south of Cape St Vincent (Fig. 6).

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THE FORMATION OF THE TEMPERATURE-SALINITY MAXIMA IN THE OUTFLOW The hydrographic surveys have demonstrated quite convincingly that the double, and perhaps triple maxima, are formed very near the Strait and, therefore, it is now more readily accepted that the bottom topography in the

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Fig. 6.—Distribution of temperature and salinity along a section running south

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from Cape St Vincent (after Ambar & Howe, 1979a).

Gulf of Cadiz can no longer be regarded as being wholly responsible for the inhomogeneity in the structure of the outflow. As well as the evidence from the Shackleton and Challenger data, this hypothesis is also supported by Zenk’s (1975a) interpretation of some observations that were made very near to Section III (Fig. 1), in which he represented the outflow in terms of three maxima. Zenk’s schematic outline of these different water types is shown in

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Fig. 7.—Distribution of temperature and salinity in the outflow (a) as it leaves the Strait at about longitude 6°18′ W and (b) through a section near III in Fig. 1 at about 6°55′ W; (after Ambar & Howe, 1979a).

Figure 8 where the properties of Layers 1 (salinity=37.50‰, density= 28.25) and 3 (salinity=36.90‰, density=27.60) can be related to the corresponding layers in Figure 7b of the Challenger data. Whereas Ambar &

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Fig. 8.—Schematic section through the Mediterranean outflow near III in Fig. 1 showing three different water types (1, 2, and 3) (after Zenk, 1975a).

Howe (1979a) chose to subdivide Layer 3, Zenk showed by means of repeated CTD casts (conductivity-temperature-depth) through Layers 1 and 2 that in this particular section Layer 2 appeared in the temperature-salinity diagrams as a discernible maximum which remained isolated from a highly variable mixing part of the water column. In Figure 9 the more stable core of Layer 2 is at a depth of 550 m with a salinity of 36.15‰ and a density of 27.45.

Fig. 9.—T-S diagram of repeated CTD records through the centre of the outflow in a section near III in Fig. 1 (after Zenk, 1975a).

Even although the interpretations of these different sets of observations have not been entirely consistent, the complexity of the structure of the outflow is now better appreciated, and an explanation of the origin of the multiple maxima will require a more comprehensive model than that suggested by Siedler (1968). The inhomogeneity in the undercurrent can probably be attributed to mixing processes occurring within the Strait of Gibraltar, and it is now known that these can be influenced by a number of short term factors, as well as the more obvious large scale seasonal changes. Defant (1961) has discussed the tidal response and the internal oscillations that take place in the transition boundary between the inflowing surface Atlantic water and the Mediterranean outflow. Significant variations in the rate of the Atlantic inflow have been attributed by Lacombe (1961) and Crepon (1965) to changes in the mean atmospheric pressure over the western Mediterranean. There have been numerous measurements of internal wave activity, either of semi-diurnal period (Frassetto, 1960; Lacombe, Tchernia, Richez & Gamberoni, 1964) or short period (Ziegenbein, 1969, 1970). Boyce (1975) has included some of these results in a theoretical discussion of the two layer flow regime within the Strait in which

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he advocates the frequent presence of steepening long internal waves and internal shock waves. Apart from this kind of internal activity, another possible mixing process was revealed after Ambar & Howe (1979a) had analysed a time series of temperature and salinity fluctuations in the boundary layer between the North Atlantic Central Water and the Mediterranean outflow in the section represented in Figure 7b. By applying the criteria outlined by Pingree (1972), which compares the slope of the regression line that is obtained by correlating the temperature and salinity fluctuations at fixed levels with the ratio of the vertical gradients of the mean temperature and salinity at the same level, they concluded that since these quantities were in good agreement then the fluctuations must be the result of either internal wave activity or vertical mixing. The time series was dominated by a semi-diurnal period, and it was recorded in a flow regime in which Zenk (1975b) has reported a mean monthly speed of 68 cm·s−1 and a significant 12.4-h period in both the velocity and temperature records. In view of the obvious slope of the surface of discontinuity between the water masses in this section, Ambar & Howe (1979a) considered an application of the Margules formula:

which, for relevant values of

and

gave in this case

. h/L is the gradient of the slope, i.e. ratio of vertical height (h) to the horizontal distance (L); f is the Coriolis factor; g is acceleration due to gravity; p and p′ are the densities of upper and lower water masses, respectively; and U and U′ are the mean velocities of the upper and lower water masses, respectively. In Figure 7b, with L=20 km, a change in h of 20 m would be easily achieved by a change in the differential velocity of 10 cm·s−1 and because of the large vertical temperature and salinity gradients a ∆h of 20 m would then readily account for the range of the semi-diurnal oscillation that was observed in the time series. The discussion then turned to the possibility that such a mixing or entrainment process might also occur within the Strait where there is a similar slope of the isohalines (Fig. 7a). Here again a depression of the boundary surface, which may be induced by quite reasonable changes in the differential velocity, will allow shallower and warmer water to be entrained and then released as part of the normal outflow. There are, therefore, a number of modulation processes that may be responsible for the double, or perhaps a multiplicity of temperature-salinity maxima in the structure of the undercurrent. Recent observations certainly testify to the degree of inhomogeneity but as yet there is not sufficient evidence to indicate which of the dynamical effects may be the most important.

MIXING IN THE OUTFLOW In a theoretical study of several overflow situations Smith (1975) developed a stream-

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tube model for bottom boundary currents which he was able to apply to the Mediterranean outflow by making use of Madelain’s (1970) data. In general, Smith concludes that the dynamics of this undercurrent would be mainly affected by friction near the source, and by entrainment further downstream. Furthermore, the external stratification and entrainment would limit its descent to 1200 m and, when clearing the continental slope, the volumetric transport will have increased from about 1.0×106m3·s−1 to 10×l06 m3.s−1. From an observational analysis of small scale thermohaline structures in the deep ocean, Pingree (1972) believes that the outflow must have mixed quite vigorously to depths of at least 1600 m on the continental slopes in the Gulf of Cadiz before spreading horizontally along isopycnal surfaces into the open Atlantic. Thereafter evidence of much deeper mixing to depths as great as 3000 m has been provided by Worthington & Wright (1970). Any attempts to quantify the rate of mixing of the undercurrent within the Gulf of Cadiz have been rare. Zenk (1975b) assumed an initial volumetric transport of 1.0×106 m3·s−1 leaving the Strait, with a water type salinity value of 38.4‰, which would then proceed to mix with North Atlantic Central Water of salinity type 35.6‰ along the normal mixing lines that are usually represented on a temperature-salinity diagram. In a two-box assessment of the cascade water budget Zenk estimated that there would be a transport of 1.75×106 m3·s−1 through Section III (Fig. 1), and that this would be composed of 54% Mediterranean water and 46% Atlantic water. As the outflow moved westward, either by following a northern route around the Spanish shelf or by flowing down several southwesterly submarine channels (Fig. 12), he calculated that this transport of 1.75×106 m3·s−1 would eventually contribute 60% to the product of a mixture with Atlantic water (40%) to produce a total outflow from the Gulf of Cadiz of 2.92×106 m3·s−1. From direct current measurements and observations of the salinity distribution in these channels, Zenk, however, calculated a total outflow transport of 2.03×106 m3·s−1. He attributed the difference to an inadequate knowledge of the mean current directions and the salinity distribution, as well as the likelihood of losses by other routes which were not monitored. By applying a quite different technique Ambar & Howe (1979a), having established the preferred flow paths of the two main Mediterranean cores, made separate estimates of the mixing rates of each of these water masses. The upper core was assumed to be a mixture of Mediterranean Water and the North Atlantic Central Water which occupies the depths between 200 m and 900 m. The main lower core was regarded as an admixture of Mediterranean water, North Atlantic Central Water and North Atlantic Deep Water which occupies depths between 400 m and 1500 m. It was necessary to decide on the relevant indices to represent these water types before the appropriate “mixing triangles” could be constructed on a temperature-salinity space. The theoretical implications and application of this technique have been fully discussed by Mamayev (1975). After a detailed analysis of their data Ambar & Howe (1979a) designated specific values to the particular sources of Mediterranean water in the outflow that supplies the two cores. Their source water type for the upper core was allocated a salinity of 37.18‰ and a potential temperature of 14.02°C with corresponding values of 38.12‰ and 13.35°C for the lower. These can be compared with Zenk’s (1970) values of 37.2‰, 13.0 °C, and 38.2‰, 13 °C, respectively which he used in his original model to account for the double maxima. The percentage

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value for the amount of Mediterranean water that was associated with each core were then calculated independently from the appropriate mixing triangles. Figure 10 shows a comparison of the results that were obtained in a section coincident with III (Fig. 1) and in a section that was made to the south of Cape St Vincent. The maximum percentage of Mediterranean water in the cores was, of course, deduced from the actual temperaturesalinity maxima that were observed in each section, and the general distribution of these values

Fig. 10.—Percentages of the Mediterranean water in the upper (Mu) and lower (Ml) cores in section (a) along III in Fig. 1 and (b) south of Cape St Vincent (after Ambar & Howe, 1979a).

throughout the region is represented by Figure 1.1. The authors concluded from this, that of the two cores the lower appears, to have mixed more vigorously in its descent from the Strait, because at this initial stage it only constitutes 70% of the mixture compared with a value of 90% for the upper core. Thereafter there seems to be a similar rate of mixing in

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both layers of the outflow as it proceeds towards Cape St Vincent. Finally, Ambar & Howe (1979b), without the benefit of any direct current measurements, deduced from geostrophic considerations a volumetric transport of 1.51×106 m3·s−1 for the undercurrent as it leaves the Strait. This is in reasonable agreement with other estimates of the outflow which include 1.65×106 m3·s−1 (Defant, 1961), 0.66 to 1.38×106 m3.s−1 (Lacombe et al., 1964), 0.72 to 1.57×106 m3·s−1 (Lacombe, 1971), 1.60×106 m3·s−1 (Bethoux, 1979) and the previously mentioned value of 1.0×106 m3·s−1 (Zenk, 1975b) which he used in his assessment of the cascade water budget. At about longitude 7°30′ W the transport had increased to 2.74×106 m3·s−1. Although this trend appears to have been reversed in the section near Cape St Vincent (2.60×106 m3·s−1), it was noted that the estimates in this area would be subject to much larger errors due to the tendency for an offshore divergence in the flow, which would not

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Fig. 11.—Isopleths of the maximum percentages in the (a) upper (Mu) and (b) lower (Ml) cores as they were observed through the Gulf of Cadiz (after Ambar & Howe 1979a).

have been monitored by the observations. Near the source, an estimate of the entrainment factor, based on the concept of mass continuity in the downstream direction, gave values between 0.02 km and 0.05 km and these were comparable with the predicted optimum value of 0.05 km from Smith’s (1975) stream-tube model.

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CURRENT VELOCITIES IN THE OUTFLOW Prior to 1970 there had been very few successful attempts to record the current speeds in the outflow and invariably these were of short duration, usually lasting just one or two days. The most relevant to this discussion were two sets of measurements, using quite different techniques, which were made at the opposite extremes of the region. Bøyum (1963) has reported results from three current meter moorings that were deployed near the Strait between 6°20′ W and 6°45′ W. Here he recorded mean speeds of 80–90 cm·s−1 over a 24-h period. In contrast, Swallow (1969), who was working off Cape St Vincent with neutrally buoyant floats, measured a westward flow of 20–30 cm·s−1 during a 2-day experiment near 36°30′ N:8°45′ W. This was associated with the main high salinity core at a depth of 1080 m. Furher offshore and centred at about 36°20′ N, he observed over a period of 7 days an anticlockwise gyre at a depth of 1400 m with speeds of 10–20 cm·s−1. This interesting effect occurred well below the salinity maximum. Even further south at 35°47′ N, and after two days of tracking, the speed at a depth of 1260 m was estimated to be 2.4 cm.s−1 towards 130 °T. It, therefore, seems likely that there may be a high degree of variability in the flow pattern of the intermediate waters in this area. More recently Madelain’s (1970) current meter measurements, although widely distributed (Stations C1–C10 in Fig. 1), were also of short duration, usually lasting just 24 h or even less. Near the Strait between 6°20′ W and 6°45′ W speeds of 100 cm.s−1 were recorded, and a substantial flow rate was maintained as far as Section III where the velocity was 50 cm·s−1. Further west the current meter moorings were positioned in the submarine canyons along Sections IV and V where there was a southwesterly flow of about 25 cm·s−1, but at the offshore mooring (36° N) this had decreased to <10 cm·s−1 with variable direction. In a 12-h period of measurements at a depth of 860 m between Sections V and VI Madelain, however, discovered a strong northwesterly flow, with speeds of 50 cm·s−1, which was to decrease to about 20 cm·s−1 at a depth of 850 m (10 m above the bottom) at the inshore current meter Station C1 in Section VII (8°30′ W). In contrast, the velocities at the southern extremity of his Section VII (36° N) were very low, usually being <4 cm·s−1. Some interesting indirect evidence of the varying speed of the outflow has also emerged from detailed studies of the microphysiographic features that have been produced by the interaction of the undercurrent with the sea bed. Heezen & Johnson (1969) identified various sea-floor zones within the Gulf of Cadiz which they have labelled as rocky, current swept bottom, sediment waves, and rolling smooth terrain, and these topographical features have been related to the gradual decrease in the speed of the undercurrent. An example of this interaction was reported by Thorpe (1972) when the presence of a suspended sediment cloud was noted beneath the main temperature-salinity maximum at about 8° W. This was assumed to have occurred in the more turbulent and fast flowing part of the outflow when it was in contact with the bottom nearer the Strait. Further evidence of these interactive effects was provided by Kenyon & Belderson (1973) after they had made a survey with the more refined side-scan sonar. The zones of scour hollows, sand ribbons, and sand waves, with related deposits, were assumed to be associated with a decreasing current flow that was still in contact with the sea floor,

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whereas in the areas of large muddy sediment waves the undercurrent was considered to be only intermittently in contact with the bottom. There was, of course, no possibility of appreciating the degree of variability in the outflow from Madelain’s short series of direct current measurements or from this type of geological survey. More recently the best attempt to study the distribution of the Mediterranean undercurrent in the Gulf of Cadiz was that undertaken by the Meteor Expedition in 1971, during which seven current meter moorings were deployed (Siedler, 1972). Measurements were made over a period of one month using Aanderaa current meters together with thermistor chains, and Zenk’s (1975b) interpretation of the flow pattern is shown in Figure 12. This is based on the mean values recorded at six of the mooring positions (see Table I), and it is the same data that Zenk used in his estimates of the outflow water budget which was discussed earlier. A more detailed account of these observations is also available in the form of a data report by Zenk (1975c). In this the records from 29 current meters and five thermistor chains are presented as a set of time series that are made up from data points that were selected at 10-min intervals. As well as the mean values, the amplitude spectra for the period range 2–120 h were computed, and they showed a consistently dominant period of 12.4 h in both the temperature and velocity spectra. In addition, a set of progressive vector diagrams were produced to illustrate the predominant current directions. Zenk notes that the high speeds in Mooring 19 include an instantaneous maximum of 84 cm·s−1 which was in fact lower than the maximum velocities of 100 cm·s−1 recorded by Thorpe (1972) at a position near to the head of the canyon and further to the west. This is attributed to the fact that the lower current meter, although only 15m above the sea bed, was still not deep enough to monitor the thin core that constitutes the undercurrent at this stage of the outflow. The main problem according to Zenk was to decide what was happening in the triangular zone between Moorings 20, 21 and 22. Mooring 20 showed the same characteristics as Mooring 19 with the maximum speeds near the bottom at a depth of 1000 m. On the other hand, at Moorings 21 and 22 the maximum speeds were recorded nearer a depth of 850 m (Table I). Although Zenk did not make any distinction, it can now be considered a possibility that this shallower stream was in fact associated with the separate upper core and, if so, these measurements would be compatible with a later interpretation by Ambar & Howe (1979b) of the flow dynamics in this area. Similarly, the relatively shallow measurements at Mooring 25 are interesting in that there were substantial speeds parallel to the shelf near Cape Santa Maria with a maximum of 69 cm·s−1 at a depth of 547 m. In addition to the Meteor cruise there were at the same time some fruitful supplementary observations being made at four other moorings, and together with CTD casts they enabled Thorpe (1976) to make an assessment of the kind of variability that there might be on a more local scale. In Thorpe’s experiment three of the moorings were clustered within 12 km of each other at about 36° 12′ N:8°03′ W. Here the Mediterranean undercurrent is no longer in contact with the sea bed but is likened to a free jet in the form of a wedge shaped flow against the continental slope to the north. The mean currents were small and generally northwesterly with a maximum speed of 7.3 cm·s−1 at a depth of 1199 m. The flow below 1100 m was shown to be less stable to vertical mixing processes than that above, with Richardson numbers of about 102 and 103, respectively. During this series of measurements there was a large adjustment in the current direction

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which suddenly began to move offshore, and then over a period of several days it proceeded

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Fig. 12.—Positions of current meter moorings Nos 19–25 that were deployed by F.S.METEOR in Spring 1971 together with the hydrographic Sections A, B, and C: CTD time series observations were made at Stations 6, 65, 78, 98 and 161; (after Zenk, 1975b).

to describe an eddy-like motion, particularly in the layers above the main core. A short distance to the west a single current meter at a depth of 1384 m recorded speeds of 30cm·s−1 where previously the current had been very weak. It was then noticed that these current variations were followed by a large change in the temperature-salinity structure of the water column with the arrival of significantly denser water in the layers below 900 m.

TABLE I Mean current and temperature data from the Gulf of Cadiz obtained during the cruise of F.S.METEOR in Spring 1971 (after Zenk, 1975b).

Ref.

Location

No.

ø

λ

Date

Depth

bottom

N

W

1971

m

m

36° 05.4′

06° 48.9′

19102

Clearance

SD

SD

Speed speed Direction Temp. temp. cm·s−1

°T

°C

27 April– 19 May

552

112

13

10

351

12.2

0.19

–27 May

649

15

68

7

324

13.1

0.15

27 April– 27 May

557

313

5

7

–

11.6

0.48

2

649

221

23

12

281

12.6

0.33

3

752

118

29

11

266

13.0

0.15

4

855

15

50

8

243

13.1

0.08

456

218

40

7

252

13.5

0.49

2

547

127

39

12

259

13.8

0.15

4

616

58

25

10

259

13.6

0.13

5

659

15

20

9

264

13.5

0.17

697

320

10

6

231

10.7

0.35

3

796

221

12

7

231

11.2

0.35

4

901

116

30

15

229

11.6

0.27

5

1002

15

44

13

236

12.6

0.16

635

465

13

7

270

12.0

0.49

5 21101

25101

20101

22101

36° 18.1′

36° 48.7′

36° 10.5′

36°

07° 18.1′

07° 55.6′

07° 27.1′

07°

29 April– 26 May

28 April– 27 May

29 April–

Oceanography and marine biology

19.0′

46.2′

75

26 May

3

884

216

39

11

272

13.0

0.13

4

984

116

22

12

274

12.8

0.13

5

1085

15

9

9

272

12.6

0.09

448

212

2

3

–

11.9

0.24

2

539

121

5

4

325

11.6

0.40

3

592

68

22

7

334

12.5

0.36

4

645

15

36

6

319

13.2

0.19

23101

36° 28.0′

07° 31.0′

28 April– 26 May

Interestingly Thorpe was able to draw attention to a similar event which occurred at another current meter station nearer to the Strait. The mooring was situated close to the head of the canyon at 36°17′ N:7°20′ W in Section IV (Fig. 1). During a 46-h period a mean current of 79.8 cm·s−1 was recorded 45 m above the sea bed and this was directed almost exactly down the slope of the canyon along 233 °T. Similarly, the speeds at 21 m and 1.8 m above the bottom were 68.9 cm·s−1 and 60.0 cm·s−1, respectively. Quite

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Fig. 13.—Temperature and salinity distribution in a section between IV and V (Fig. 1) with salinity values above 30‰ (after Ambar & Howe, 1979a).

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suddenly there was again a considerable increase in both the salinity and the thickness of the Mediterranean water, which amounted to a change in the potential density values of the deepest layer from 27.79 to 27.97. Although it is possible that these two events were connected, the reasons for such variations in the flow regime could not be conclusively explained and it was, therefore, tentatively suggested that certain processes of modulation might be affecting the undercurrent as it leaves the Strait. The only other attempt to determine the velocity field in the outflow has been made by Ambar & Howe (1979b) who computed the geostrophic components of the flow through the numerous temperature-salinity sections

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Fig. 14.—Velocity field (cm·s−1) in a section near IV (Fig. 1) calculated from the T-S distribution shown in Fig. 13: the alignment of the stations in this section is shown (inset) so that positive directions of flow will be 306° T for Stations 55–58, and 355° T for Stations 50–55; (after Ambar & Howe, 1979b).

that were available from the Challenger 1976 data. Their calculations were based on an

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assumed level of no motion between depths of 400 and 600 m, and in several areas they found that the results were in good agreement with the direct measurements of Madelain (1970) and Zenk (1975c). The maximum estimated speed of the outflow leaving the Strait was 180 cm·s−1 at a depth of 345 m. After the separation at about 6°40′ W the upper core was allocated an initial velocity of 46cm.s−1 as it proceeded around the Spanish continental slope. In the more complex region near 7°30′ W the temperature-salinity sections became less uniform (Fig. 13) and consequently the interpretation of the associated velocity field became more demanding

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Fig. 15.—Flow paths of the upper (a) and deep (b) cores of the outflow: maximum velocities are shown (cm.s−1) and in (b) the presence of a deep countercurrent is represented by the open arrows; after Ambar & Howe, 1979b).

(Fig. 14). For example, Ambar & Howe concluded that Figures 13 and 14 represent a meandering in the flow. By relating the computed maximum velocities to the temperature-salinity maxima in the two cores the preferred flow paths were deduced and plotted as shown in Figure 15. This separation, together with the subsequent meandering of the main lower core at about 7°30′ W would be fairly compatible with Zenk’s (1975b) presentation of the flow pattern, and might also account for some of the variability that was observed by Thorpe (1976). In the latter case the changes in speed and physical properties of the outflow could have been caused by a lateral displacement of the undercurrent which is quite feasible in such a meandering regime. This might then induce an eddy type motion in the water column and easily explain the large changes in the density at certain depths. As a final comment the Ambar & Howe (1979b) geostrophic analysis produced one further interesting result which concerns the deep circulation in the region south of Cape St Vincent (Fig. 15). Here the velocity profiles indicated the presence of a deep counter current beneath the main core of the Mediterranean water. In the six sections that were made between 8°30′ W and 9°30′ W the estimated speeds at depths of 1400 to 1500 m varied between 6 and 22 cm·s−1. Apart from the eddy motion that was described by Swallow (1969) there have been no other reports of any unusual effects in the deep circulation of this area and, as yet, there is no direct confirmation of such a counterflow. Indeed the current meter data which Zenk (1975c) has produced for Mooring 24 (Fig. 12) shows that at a depth of 1568 m the mean monthly speed was 3.2 cm.s−1 towards 272 °T. After a more detailed inspection of the low passed filtered data Zenk (1980) has, however, claimed that there is a definite tendency for the flow to reverse for a period of one or two days with speeds approaching 10 cm.s−l.

REFERENCES Ambar, I., Howe, M.R. & Abdullah, M.I., 1976. Dt. Hydrogr. Z., 29, 58–68. Ambar, I. & Howe, M.R., 1979a. Deep-Sea Res., 26A, 535–554. Ambar, I. & Howe, M.R., 1979b. Deep-Sea Res., 26A, 558–568. Bethoux, J.P., 1979. Oceanol. Acta, 2, 157–163. Boyce, F.M., 1975. Deep-Sea Res., 22, 597–610. Bøyum, G., 1963. Bergen, Geophysical Institute. (NATO sub-committee on oceanographic research), 27 pp. Crepon, M., 1965. Cah. océanogr., 17, 15–32. Defant, A., 1936. Deutsche Atlantische Expedition, “Meteor” 1925–1927, Band VI— Atlas, Verlag von Walter de Gruyter and Co., Berlin, 103 plates. Defant, A., 1961. Physical Oceanography. Vol. 1, Pergamon Press, Oxford, 729 pp. Frassetto, R., 1960. Deep-Sea Res., 7, 152–162. Gieskes, J.M., Meincke, J. & Wenk, A., 1970. “Meteor” Forsch-Ergebn. A., 8, 1–11.

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Heezen, B.C. & Johnson, G.L., 1969. Bull. Inst. océanogr. Monaco, 67, No. 1382, 95 pp. Howe, M.R. & Tait, R.I., 1972. Deep-Sea Res. 19, 781–791. Howe, M.R., Abdullah, M.I. & Deetae, S., 1974. J. mar. Res. 32, 377–386. Kenyon, N.H. & Belderson, R.H., 1973. Sediment. Geol., 9, 77–99. Lacombe, H., 1961. Cah. océanogr., 13, 73–107. Lacombe, H., 1971. Notes and Mem. Serv. Geol. Maroc, 222, 111–146. Lacombe, H., Madalain, F. & Gascard, J.C., 1968. Cah. océanogr., 20, 101–108. Lacombe, H., Tchernia, P., Richez, C. & Gamberoni, L., 1964. Cah. océanogr., 16, 283– 314. Madelain, F., 1970. Cah. océanogr., 22, 43–61. Mamayev, O. L, 1975. Temperature—Salinity Analysis of World Ocean Waters. Elsevier Sci. Publ. Co., 374 pp. Pingree, R.D., 1972. Deep-Sea Res., 19, 549–561. Siedler, G., 1968. Kieler Meeresforsch., 24, 59–65. Siedler, G., 1972. “Meteor” Forsch-Ergebn. A, 10, 79–95. Smith, P.C., 1975. Deep-Sea Res., 22, 853–873. Swallow, J.C., 1969. Deep-Sea Res., Suppl. to 16, 285–295. Thorpe, S.A., 1972. Nature, Land., 239, 326–327. Thorpe, S.A., 1976. Deep-Sea Res., 23, 711–727. Worthington, L.V. & Wright, W.R., 1970. North Atlantic Ocean Atlas, Woods Hole Oceanographic Atlas Series Vol. II, Woods Hole, Mass., U.S.A. Zenk, W., 1970. Deep-Sea Res., 17, 627–631. Zenk, W., 1975a. “Meteor” Forsch-Ergebn. A, 16, 35–43. Zenk, W., 1975b. “Meteor” Forsch-Ergebn. A, 16, 23–34. Zenk, W., 1975c. “Meteor” Forsch-Ergebn. A, 15, 20–48. Zenk, W., 1980. Deep-Sea Res., 27A, 97–98. Ziegenbein, J., 1969. Deep-Sea Res., 16, 479–487. Ziegenbein, J., 1970. Deep-Sea Res., 17, 867–876.

THE RÔLE OF BACTERIA IN THE TURNOVER OF ORGANIC MATTER IN THE SEA I.R.JOINT Natural Environment Research Council, Institute for Marine Environmental Research, The Hoe, Plymouth, U.K. and R.J.MORRIS Institute of Oceanographic Sciences, Wormley, Godalming, Surrey, U.K.

Oceanogr. Mar. Biol. Ann. Rev.. 1982, 20, 65–118 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION The study of bacteria in the sea has two historical origins. The first is classical bacteriology which developed from the work and philosophy of Robert Koch and involves isolating a bacterium, growing it in axenic culture and identifying it by a series of determinative tests. The second origin is the more recent approach of marine biologists who have studied the rate of a process such as primary production, without being particularly concerned with identifying the organism involved. The classical bacteriological approach was used in early studies of marine bacteria but it is arguable whether any significant advance was made in our understanding of the rôle of bacteria in the sea until research began on the rates of bacterial activity. The basic problem with the classical bacteriological approach is that any interaction between species is destroyed by isolating a bacterium and growing it in culture. It is also impossible to know whether the isolated bacterium is representative of the population from which it was derived; it is all too easy, using an appropriate selective medium, to isolate almost any type of bacterium from coastal water or sediments. It is, however, not enough to show that the isolate is capable of a particular reaction; the rate of processes involving bacterial activity must be measured in situ before it is possible to understand the ecological significance of a bacterial species. The early obsession with culturing single bacterial species may have delayed our understanding of many important processes mediated by bacteria because the act of isolation destroys any synergistic reactions occurring between species. The reconstruction of interspecies interactions using pure cultures has a rôle and has been used successfully, for example, in elucidating interactions between fermentative bacteria in sediments which

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result in the complete mineralization of organic matter. It is, however, the

exception, rather than the rule, that marine bacteriologists consider commensalism as important in the turnover of organic matter by marine bacteria; Robert Koch’s philosophy still pervades microbial ecological research. The purpose of this review is to consider the importance of bacteria in the turnover of organic matter in the sea. The intention is to consider only autochthonously produced organic matter and the importance of bacteria in utilizing anthropogenic waste matter in estuaries and coastal waters will not be considered; inevitably, more research has been done in these areas than the deep sea, so examples will be taken from estuaries and coastal waters when no other data are available. The productivity, and hence autochthonous supply of organic matter, varies considerably in different oceans and regions, with conditions ranging from excess organic matter in upwelling regions to very low concentrations of organic matter available for marine bacteria in oligotrophic regions. It is, clearly, impossible to draw many generalized conclusions on bacterial activity from such diverse environments and to discuss the activity of marine bacteria in general would be unwise. The problems, however, encountered by marine bacteriologists in studying such diverse environments are often the same and the common theme in all bacterial studies must be, how important are bacteria to the productivity of the seas? The review will begin by considering the sources of supply of organic matter and their availability to marine bacteria; the problems involved in measuring bacterial biomass and the effect of environmental conditions on bacterial activity will then be considered.

SUPPLY OF ORGANIC MATTER Bacteria are incapable of ingesting particulate organic matter; they have no mechanism of phagocytosis or pinocytosis and all organic matter must be soluble before it can be transported across the cell membrane. Of course, bacteria do utilize particulate organic matter but these complex organic molecules must be acted upon by extracellular enzymes to produce soluble organic compounds. Dissolved organic matter can, therefore, be considered as the preferred substrate for bacteria since it can be utilized without any investment of energy for the production of extracellular enzymes. Not all dissolved organic matter can, however, be utilized by bacteria and the great part of the organic matter in the sea must be considered as biologically refractory and not available for microbial growth (Jørgensen, 1976). It is perhaps worth considering the definition of dissolved matter. The practical definition, which is accepted by most marine chemists is that organic matter which passes through 0.45 µm pore size filter paper is considered as “dissolved” and that which is retained, is “particulate”. Clearly, this definition of dissolved organic matter is not adequate when considering bacterial metabolism because the largest molecules that can be transported across the bacterial cell envelope have molecular weights of a few thousand Daltons. Indeed the marine chemists’ definition of dissolved organic matter actually includes many particles, some of which are bacteria. So, any data cited on dissolved organic matter cannot be assumed to include organic matter which bacteria can

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directly assimilate. In the absence, however, of any measurements which a bacteriologist would accept as representing dissolved organic matter, we must rely on the chemists’ working definition. DISSOLVED ORGANIC CARBON Williams (1975) has reviewed the sources of supply of organic matter to the oceans and has summarized reported values for dissolved organic carbon (DOC) and the typical variations that occur in DOC concentrations. Williams (1975) assumed a mean DOC concentration of 0.7 mg C·1−1 and estimated that the total DOC reservoir in the sea was 1×1018 g C; much of this carbon is probably biologically refractory which accounts for the constancy of this figure in different regions of the sea. The sources of DOC are the atmosphere (Szekielda, 1978), rivers, and production in situ; Williams (1971) estimated that precipitation brings organic matter equivalent to 1% of the total net production to the oceans and Williams (1975) estimated that rivers supply a similar amount. So by far the largest contribution to DOC in sea water is produced in situ. CHEMICAL SPECIATION OF ORGANIC COMPOUNDS IN THE SEA Bacteriologists would like to know something about the chemical composition of naturally occurring organic matter so that they can experiment on the growth efficiencies of marine bacteria on natural substrates. Relatively few of the organic compounds that microbiologists would consider as important and central to the metabolism of bacteria, have been measured in the sea. Obviously the organic composition of marine organic matter is extremely complex and marine microbiologists would be naïve to expect marine chemists to produce a complete description of the chemical composition. Moreover, simple, rapid procedures are not readily available for measuring such centrally important compounds as acetate in the sea and the very low concentrations of biologically labile substances increase the difficulties in devising analytical procedures. It is obviously easier to study compounds which are present at relatively high concentrations but, almost by definition, such compounds are unlikely to be important in the cycling of organic matter because their high concentrations are a result of non-utilization by bacteria. Some data are available on certain classes of organic compounds such as lipids (Morris & Culkin, 1975; Morris & Eglinton, 1977 and references therein), carbohydrates (Burney & Sieburth, 1977; Liebezeit, Bölter, Brown & Dawson, 1980; Ittekkot, Brockmann, Michaelis & Degens, 1981) and amino acids (Williams, 1975). More recently, Daumas (1976) has studied the relationship between dissolved amino acids and the protein fraction of particulate organic matter (POM) in the Mediterranean and has found that combined amino acids (peptides) were always more abundant than free, dissolved amino acids. Data on specific carbohydrates and amino acids in a variety of deep-sea sediments are given by Degens & Mopper (1976) and in shelf seas by Litchfield, Munro, Massie & Floodgate (1974); recently Miller, Brown, Pearson & Stanley (1979) have quantified some low molecular weight organic acids in a sediment polluted with pulp-mill waste.

TABLE I

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Primary production and excretion of organic matter by phytoplankton as determined by the 14C method

Region

Particulate production mg C·m3.h−1

Dissolved production mg C·m−3·h−1

% excretion

Reference

1. Coastal N. Pacific

32.1–184

3.6–15

7–11 Anderson & Zeutschel (1970)

2. Offshore N. Pacific

6–12.25

1.2–2.2

13–26 Anderson & Zeutschel (1970)

3. U.S.A. estuaries

27–38

0.08–2.4

0.2–6.8 Thomas (1971)

4. Coastal Atlantic

0.4–18.5

0–0.57

0–27.4 Thomas (1971)

5. N. Atlantic

6.9–20.7°

3.1–32.6a

6. Coastal N. Pacific

6.6–176ab

1.75–19.5ab

7. Coastal subtropical Atlantic

0.14–1.05b

0.003–0.11b

10–208a

0.61–18.16a

<1–37

<0.2–2.55

8. N.W. Africa 9. Coastal Atlantic aData expressed as

21–69.7 Choi (1972) 9.8–26.9 Berman & HolmHansen (1974) 1–23 Williams & Yentsch (1976) 1.3–22.3 Smith et al. (1977) 5–33 Mague et al. (1980)

mg C·m−2·h−1 for depth of euphotic zone. rate assuming 12 h daylight.

bCalculated from daily

Studies such as these are contributing towards knowledge of naturally occurring compounds which are likely to be the preferred substrates of marine bacteria but there is a need for greater collaboration between chemists and bacteriologists. PHYTOPLANKTON PRODUCTION OF DISSOLVED ORGANIC CARBON The fraction of phytoplankton production most readily available to marine bacteria is soluble organic matter excreted by photosynthetic phytoplankton. There has been considerable controversy in the literature on the magnitude of this excretion with reports varying from less than 1% to 70% of phytoplankton production (Table I). Many of the very high percentages should be treated with caution because they were often the result of measurement close to the limits of detection of the method and the errors involved in the estimations must be very large. Other experimental artefacts have been implicated and Sharp (1977) has reviewed the problems of estimating the flux of dissolved organic carbon from phytoplankton. In a recent comprehensive study of DOC excretion by

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natural phytoplankton populations, Mague, Friberg, Hughes & Morris (1980) concluded that extracellular release did occur but that it was a minor component of total primary production in coastal waters (<5 mg C·m−3·h−1), with most of the light energy captured by phytoplankton going into biomass production. Although DOC excretion may not be a very large proportion of the carbon budget of a phytoplankton cell, it may, however, be an important source for bacteria, especially when one considers that the organic compounds excreted by phytoplankton are small molecular weight intermediates of metabolism which can be taken up and utilized by bacteria. PHYTOPLANKTON PRODUCTION OF PARTICULATE ORGANIC CARBON Phytoplankton biomass has usually been assumed to be completely consumed by zooplankton grazing (Steele, 1974) so, at first sight, phytoplankton cells are not a direct source of organic matter for bacteria. There is, however, evidence that in certain seasons, there is a mismatch in zooplankton biomass and phytoplankton production with insufficient zooplankton to utilize all of the phytoplankton production (Colebrook, 1979). Lindley & Williams (1980) suggested that the zooplankton biomass in the Fladen Ground of the North Sea was insufficient to control the spring phytoplankton outburst; in a subsequent paper, however, Williams & Lindley (1980) reversed this view and speculated that there was insufficient phytoplankton production to sustain the observed zooplankton biomass. The evidence is, therefore, confused but it is worth considering that not all of the phytoplankton production is consumed directly by zooplankton and that a proportion may be available for bacterial production. Further evidence comes from sediments underlying highly productive shallow waters, especially of upwelling regions, where organic matter appears to be the result of direct input of dead or dying phytoplankton cells, rather than a result of secondary production (Morris & Calvert, 1977). PRODUCTION OF ORGANIC MATTER BY ZOOPLANKTON A significant proportion of the phytoplankton ingested by zooplankton may also be released as POC and DOC. Watson (1978) argued that the efficiency of utilization of ingested food at each trophic level was only 50–60% and that 10–20% was required for growth with another 30–40% for maintenance. Therefore, at each trophic level, 40–50% of the ingested food is excreted as DOC or released as POC in faeces, and hence a very large proportion of the primary production would seem to be available for bacterial utilization. There have been very few studies which have attempted to measure the proportion of ingested organic matter which is released as DOC. Webb & Johannes (1967) measured the release of amino acids by net zooplankton and estimated that an amount of amino acid equal to the concentration in sea water was released by zooplankton in one month. Dagg (1974) studied the efficiency of carnivorous feeding and found a significant proportion (>25%) of the prey contents was lost before ingestion. In a study of freshwater zooplankton, Lampert (1978) found that up to 10% of the algal carbon removed by

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grazing was lost as DOC, both from algae damaged during feeding and by excretion from the animals and leaching from their faeces. In a laboratory experiment with Calanus pacificus, using 14C as a tracer, Copping & Lorenzen (1980) found that 27% of the grazed phytoplankton carbon appeared as DOC, 24% as dissolved inorganic carbon from zooplankton respiration, 3–4% as faecal pellets, and 45% was incorporated into the copepod body. These data can be used to construct a hypothetical food web which incorporates the ideas of Watson (1978) on trophic efficiency. Figure 1a shows the “classical food-web” as described by Steele (1974) for the North Sea; data were not available for bacteria hence their inclusion in brackets and Steele also assumed that microbial degradation of zooplankton faeces occurred on the bottom in shelf areas. The result of this food web is 2% of phytoplankton production is found in pelagic carnivores and 33% is transported to the benthos as faeces. Figure 1b attempts to include DOC into a pelagic food web. It is assumed that 10% of primary production is directly excreted by phytoplankton (Table I) and that all of the phytoplankton is eaten by herbivorous zooplankton. The data of Copping & Lorenzen (1980) suggest that 27% of ingested food is released as DOC (i.e. 24% of total phytoplankton production). All herbivores are then assumed to be eaten; the data of Dagg (1974) suggest that feeding of carnivores is much less efficient than that of herbivores, so DOC production is taken to be 50% of grazed food (i.e. 25% as loss of prey on capture and 25% by the same excretion processes as herbivores). So a two stage grazing chain results in the loss of 54% of primary production as DOC and 5% loss as POC through faeces production. If it is assumed that 90% of this organic matter does not enter the pool of refractory organic matter in the sea but is utilized by bacteria with high efficiency (see Table III and p. 85), then 40% of the annual phytoplankton production could end up as bacterial biomass. Clearly such mass balance calculations are of limited value in determining the true trophic rôle of bacteria because of the assumptions made in constructing the budget. A direct result of inefficient zooplankton grazing is,

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Fig. 1.—a, part of the food web of the North Sea (redrawn from Steele, 1974). b, hypothetical pelagic food web showing the proportion of organic matter produced as DOC (——) and POC (----) at each trophic level using herbivore grazing efficiencies from Copping & Lorenzen (1980) and carnivore efficiencies from Dagg (1974): numbers in brackets indicate the amount of primary production which passes through each trophic level or individual process and the “earth (ground)” symbol indicates loss of carbon by respiration or to the refractory DOC pool.

however, the production of considerable amounts of DOC. The contribution of zooplankton-derived DOC to the total supply of DOC in the sea is a direct result of the amount of phytoplankton production eaten; if current dogma is correct and most phytoplankton production is grazed by pelagic herbivores, then zooplankton must be a major source of supply of dissolved organic matter which could be utilized by marine bacteria. If zooplankton do not use all the phytoplankton production, a considerable amount of organic matter will also be available to heterotrophic microbes.

VERTICAL FLUX OF ORGANIC MATTER IN THE OCEANS Most of the organic matter in the ocean results from production in situ in the euphotic zone, but this layer represents a very small fraction of the sea. It is important to know how much of this production is utilized in the euphotic zone and how much of it sinks to sediments in coastal water and to the rest of the deep water column, which may be over 5000 m in depth. SINKING RATES OF ORGANIC PARTICLES The growth of phytoplankton cells which are not grazed by zooplankton will eventually be limited by nutrient availability or by sinking out of the euphotic zone and would ultimately be available to bacteria within the water column or sediment. The sinking rates of individual phytoplankton cells are too low to contribute significant amounts of organic matter to deep-sea sediments; Smayda (1970) reported sinking rates of marine phytoplankton ranging from <0.1 m·day−1 for small living algae to>500 m·day−1 for large dead cells, Lännergren (1979) recorded a range of 0–9 m·day−1, and Bienfang (1980) found from 0.34–1.65 m·day−1 for a natural phytoplankton population. A more important route is the sinking of faecal pellets (Schrader, 1971; Knauer, Martin & Bruland, 1979; Wiebe, Boyd & Winget, 1980). In a laboratory study, Turner (1977) measured sinking rates of from 15–153 m·day−1 for faecal pellets produced by the copepod Pontella meadii, fed on four different unialgal cultures. The sinking rates of individual faecal pellets was not constant because their small volume and density made them very susceptible to variations in water column micro-structure. Fowler & Small (1972) found sinking rates for euphausiid faecal pellets of 240 m·day−1. To a certain extent, these studies are artificial because the zooplankton were fed on unialgal diets

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rather than a natural mix of phytoplankton cells but they do give an indication of the likely rates in the sea. Knauer et al. (1979) estimated the flux of faecal pellets in the northeastern Pacific to be about 1000 faecal pellets·m−2·day−1; Wiebe et al. (1980) measured a flux of 160 faecal pellets·m−2·day−1 off the Bahamas and confirmatory laboratory studies found sinking rates of 50–940 m·day−1 with a mean of 159 m·day−1. Faecal pellets, therefore, sink quickly enough to supply organic matter to deep-sea sediments but, even whilst sinking, organic matter is likely to be broken down by bacterial activity and so the amount of organic matter reaching the sediments will be reduced. Honjo & Roman (1978) studied the rate of breakdown of copepod faecal pellets and found that degradation of the surface membrane of the pellet took 3 h at 20°C but that at 5°C, the temperature of much of the deep ocean, surface membranes remained intact for up to 20 days. After degradation of the membrane, the pellet lost integrity and broke up into small amorphous aggregates. Therefore, if faecal pellets survive bacterial degradation whilst sinking through the warm surface waters, there is a good chance that they could survive to sink to the deep-sea sediment. Honjo & Roman (1978) collected faecal pellets in a particle trap at 2200 m but most of the pellets had lost their surface membrane, which the authors assumed was the result of bacterial activity in the warm surface waters. Iturriaga (1979) studied the changes that occur in sedimenting organic matter as a result of bacterial activity. At 20°C, he found that phytoplankton cells decomposed at a rate of 35%·day−1 and that “zooplankton”, as exemplified by Artemia salina, decomposed at 18%·day−1; at 5°C the rates were much lower, about 3%·day−1 for phytoplankton and 8%·day−1 for Artemia. Comparable decomposition rates were recorded by Harding (1973) who found that killed copepods decomposed within 3 days at 20°C and within 11 days at 4°C and Harding (1973) commented that copepods which die in the warm surface waters are unlikely to be recognizable as copepods if caught by a plankton net within one or two days of death. In Situ MEASUREMENT OF VERTICAL FLUX Within the last decade, a considerable body of data has been obtained on the total flux of particles in the deep sea, including the sinking of organic particles from the surface layers of the ocean through the thermocline to the marine sediment. Early estimates of particulate flux were based on the analysis of large volume water samples taken at various depths (Hobson, 1967; Gordon, 1970a, b; Feely, Sackett & Harris, 1971; CopinMontegut & Copin-Montegut, 1972; Harris, 1972; Meade et al., 1975; Krishnaswami & Sarin, 1976). It was, however, obvious that water bottle samples were not a good way of estimating flux because the relatively rare, rapidly settling particles were unlikely to be sampled. McCave (1975) showed that most of the particles in the deep ocean were small (<32 µm diameter), were not moving downwards very fast, and constituted a “background”; rarer, larger particles contributed most of the flux of organic matter but because of their scarcity, were not sampled by the 10 to 30–1 water bottles. It was only with the deployment of sediment traps (Honjo, 1978; Knauer et al., 1979; Honjo, Connell & Sachs, 1980; Jannasch, Zafiriou & Farrington, 1980; Wiebe et al., 1980) and large volume in situ filtering systems (Bishop & Edmond, 1976) that reliable information

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began to be obtained about the flux of all particles through the water column. Bishop et al. (1977) using an in situ pumped system observed that the largest concentration gradient in organic carbon, nitrogen, and phosphorus was between 50 and 100 m and below that depth, concentrations decreased very slowly with depth. Using a model for particle settling, Bishop et al. (1977) estimated that 99% of the vertical flux was due to faecal pellets although they comprised only 5% of the total particulate organic carbon (POC) concentration; faecal pellets were estimated to have a transit time of 10 to 15 days in a 4 km water column. More important, calculation showed that recycling in the upper 400 m accounted for 87% of organic carbon, 91% of the nitrogen, and 94% of the phosphorus. Similarly, Bishop, Ketten & Edmond (1978) found that 90% of the autochthonous organic matter in the euphotic zone of the southeastern Atlantic was recycled in the upper 400 m. Sampling at a deep station at 1500m in the eastern equatorial Pacific, Bishop, Collier, Ketten, & Edmond (1980) found much lower particle concentrations and fluxes than they had found in the Atlantic but this merely reflected the lower productivity of the surface waters. There are, however, still doubts about the sampling methodology employed by Bishop et al. (1980); sediment traps deployed by Cobler & Dymond (1977) for 234 days in the area studied by Bishop et al. (1980) indicated a flux 40 times greater than that estimated by large volume filtration. The traps collected material at a rate 4–4.5 mg C·m−2·yr−1 but the mass flux calculated from the depth profile data indicated only 0.1 mg C·m−2·yr−1. Bishop et al. (1980) suggested that this discrepancy might be due to a temporal nonlinearity in the proportion of surface production reaching the sea bottom; at times of high primary production, zooplankton grazing might not utilize all the phytoplankton and the increased food supply could result in the production of more faecal pellets. This would be reflected in a short-term increase in organic matter flux to the deeper water which would accumulate in sediment traps but which could not be detected in water samples unless the sampling coincided with the period of increased transport. Support for this hypothesis comes from Knauer et al. (1979) who found a high particle flux in mid-water when primary production was high in the euphotic zone. An estimate of flux obtained from instantaneous depth distribution may not, therefore, give a good estimate of the annual sedimentation of organic matter but, changes in organic matter kept in sediment traps for weeks or months, also pose problems in interpretation of data; it is arguable which method gives the more realistic estimate of sedimentation. Honjo (1978) deployed sediment traps 100 m above the sea bed in the Sargasso Sea and estimated the vertical flux to be 0.5% of the total carbon flux in the area; pigmented particles of one type of faecal pellet were the major source of organic carbon in this area. Clay was also a significant component of the material in the traps but this resulted from resuspension of bottom sediment and was not derived from surface water. Other workers have also found that a significant amount of the material collected in sediment traps is derived from resuspension of bottom sediments; Spencer et al. (1978) found the largest flux of organic matter to be due to the re-sedimentation of suspended sediment, aggregated by the action of benthopelagic organisms into faecal pellets which were 80% clay. Most of the organic matter (90%) was contributed by rapid settling of large particles derived from the surface ocean. Therefore, the data from particle traps indicate that the dominant mechanism of vertical

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transport to the deep ocean is the settling of rare, large particles such as faecal pellets (Suess, 1980) or “marine snow” (Gordon, 1970a; Silver, Shanks & Treni, 1978) of >200 µm diameter. Suess (1980) derived an empirical equation to predict the organic carbon flux at any depth in the ocean as a fraction of net primary production and depth dependent consumption. Müller & Suess (1979) have derived an empirical expression for the amount of organic matter fossilized in pelagic sediments which implies that the organic content of sediments doubles with a 10-fold increase in sedimentation rate. ESTIMATES OF ORGANIC FLUX FROM ISOTOPIC RATIOS Another method of estimating flux of organic matter in the ocean is based on stable isotope ratios of carbon and, to a lesser extent, nitrogen (Saino & Hattori, 1980; Sweeney & Kaplan, 1980) and oxygen (Craig & Kroopnick, 1970; Kroopnick, 1975). The stable isotopes of carbon, 12C and 13C, have different abundances in the atmosphere and ocean with an enrichment of about 7‰ in 13C (Deuser & Degens, 1967). In addition, the two isotopes are taken up at different rates by phytoplankton because the primary enzyme of carbon fixation, ribulose 1, 5 bisphosphate carboxylase, discriminates against 13C (Christeller, Laing & Troughton, 1976) and results in a depletion of 13C in photosynthetic organisms (Degens, Guillard, Sackett & Hellebust, 1968). Isotopic fractionation is also temperature-dependent and has been used to indicate well-defined water masses (Fontugne & Duplessy, 1978, 1980). 13C/12C ratios of marine dissolved and particulate organic carbon were first studied by Williams (1968) and Williams & Gordon (1970) who found a constant ratio with depth and time. This suggested a common source for DOC and POC and that the organic carbon was refractory and not subject to microbial transformation. Eadie & Jeffrey (1973), in a more comprehensive sampling programme, found, however, that the isotopic fractionation indicated a change in the organic composition of deep-water to sediment DOC; their data, therefore, suggested that organic matter was not biologically refractory and that the surface sediment was a region of relatively high bacterial activity. These results, however, have not been supported by subsequent work; Bishop et al. (1977) found ambiguous results which were not easy to interpret but which tended to confirm the refractory nature of organic matter in the deep sea. Williams (1968) estimated the age of DOC in the ocean using radiocarbon dating techniques and gave an apparent age before present of 3470 and 3350 years for two samples from the northeastern Pacific. Inorganic carbonate was found to be younger (2194 and 1480 years) which is confirmation of the refractory nature of most of the oceanic DOC and suggests an extremely slow turnover of organic matter. Subsequently, Williams, Stenhouse, Druffel & Koide, (1978) used the perturbation in natural 14C abundance caused by atmospheric H-bomb tests to measure the flux of organic matter to the deep-sea sediment. The results, however, were puzzling because organic matter with 13C/12C ratios which suggested a recent origin was found down to at least 12 cm in the sediment; this did not seem to be the result of bioturbation of the surface sediment since thorium isotopes showed a smooth distribution with depth of sediment. The density of animal burrows was also low and did not appear to be sufficient to account for the penetration of recent organic matter into the sediment. Williams et al., (1978), however,

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accepted that it was difficult to conceive of a mechanism other than bioturbation to explain the observed distribution; Peng, Broecker, Kipphut & Shackleton, (1977) implicated bioturbation in mixing 14C in the top 6–18 cm of sediment from the Indian Ocean. CONCLUSIONS ON THE VERTICAL FLUX OF ORGANIC MATTER It would appear that most autochthonous organic matter is recycled within the surface layers of the sea and that the proportion of primary production which sinks to the deeper waters is probably significantly <10%. The major route of organic flux to the deep ocean is by large organic particles such as faecal pellets and the bulk of the dissolved organic carbon in the deep ocean is refractory and of a great age. From a marine bacteriologist’s point of view, it is clear that most bacterial activity must occur in the surface few hundred metres of the sea, except in coastal waters, where the proximity of the sediment to the euphotic zone results in little degradation of sedimenting organic carbon in the water columns and hence, increased organic supply to the sediment.

ESTIMATION OF BACTERIAL BIOMASS Bacteriology evolved as a scientific discipline because it was possible to culture bacteria on defined medium and hence to grow large populations from a single bacterium; indeed, the basis of much bacterial taxonomy is the ability or inability to grow on a defined substrate. It is, therefore, not surprising that the earliest workers in marine bacteriology used the techniques developed by medical bacteriologists and relied almost exclusively on culture techniques to enumerate marine bacteria (ZoBell, 1946). They soon found, however, that the large body of data available for the classification of pathogenic bacteria was of little use when applied to marine isolates. The lack of determinative tests for nonpathogenic bacteria meant that it was difficult to identify and classify many of the species isolated from the sea and, consequently, the classical approach to marine bacteriology evolved very slowly. Nevertheless progress has been made in the taxonomy of marine bacteria (Hodgkiss & Shewan, 1968). There is no single method for the determination of bacterial biomass which is used universally and which does not have drawbacks; the commonest methods will be discussed here. MICROSCOPY Bacterial populations can be quantified by direct microscopic examination or by a variety of chemical procedures. The problem common to both approaches is that the abundance of bacteria in the sea is very low and some procedure is required to concentrate the cells. The commonest approach is to filter water samples through a filter which retains bacteria; the size of particle which can be retained on a filter depends on the filter type and filter characteristics (Sheldon & Sutcliffe, 1969; Sheldon, 1972). These concentration

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procedures, however, have disadvantages for direct microscopy because the opacity of the filters means that bright-field or phase-contrast microscopy cannot distinguish bacteria from filter unless the bacteria are stained. Early studies employed standard histological stains such as Imido-Schwarts (Wiebe & Pomeroy, 1972) but it was difficult to distinguish bacterial cells from detritus. Considerable advances have been made in recent years by the application of epifluorescence microscopy to the study of marine bacteria; this has proved particularly effective when coupled with fluoro-chromes such as Acridine Orange which stains DNA and RNA (Daley & Hobbie, 1975; Hobbie, Daley & Jasper, 1977; Ramsey, 1978) or, more recently, DAPI (4,6-diamidino–2-phenylindole) which binds to DNA and is reported to be a better fluorochrome than Acridine Orange for distinguishing marine bacteria from detritus (Coleman, 1980; Porter & Feig, 1980). These procedures have the disadvantage that they do not distinguish living from moribund or dead bacteria, because any nucleic acid present will bind the fluorochrome and fluoresce; other, less convenient procedures must, therefore, be used to determine viable numbers of bacteria. AUTORADIOGRAPHY One promising method, which has not been used very much, is the microautoradiographic method of Brock & Brock (1966) and Brock & Brock (1968) which can be used in combination with epifluorescence microscopy to determine the proportion of stained bacteria capable of metabolizing 14C-labelled organic substrates. The results so far obtained suggest considerable variability in bacteria activity with different substrates. Hoppe (1974) found that only 30% of a bacterial population took up 14C glucose, 9% took up 14C acetate but 60% took up 3H amino acids (Hoppe, 1976). Sieburth (1979) has criticized this technique because different bacteria have different substrate specificities and it is difficult to supply a mixture of labelled substrates which one could be sure of being taken up by all bacteria. This criticism does not, however, apply to the work of Faust & Correll (1976) who did not supply an organic substrate but used the uptake of 33P-inorganic phosphate to distinguish metabolically active bacteria from senescent or dead bacteria; they found that the proportion of active bacteria varied between 63 and 85% of the total. It, therefore, appears that a significant proportion of the bacteria stained by Acridine Orange may be moribund and are not involved in heterotrophic activity. CULTURE TECHNIQUES The development of epifluorescence microscopy to enumerate aquatic bacteria confirmed what had been suspected for many years; the culture techniques favoured by early microbiologists detected only a small proportion of bacteria present in the sea. Using phase contrast microscopy, an inferior technique to epifluorescence microscopy, Taga & Matsuda (1974) found that the number of bacteria which grew on a marine bacteriological medium varied from 17% of the total number in a coastal region to 0– 006% of the total in the Pacific Ocean. Hoppe (1976) found that plate counts detected only 0.01 to 6% of active bacteria; in addition, the majority of bacteria were very small, passing through 0.4 µm pore NucleporeR filter and these very small bacteria could not be

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cultured. The existence of this large population which defies established culturing techniques led Sieburth (1979) to propose that there are two populations of marine bacteria; autochthonous “planktobacteria” which are very small, free-living bacteria, utilizing dissolved organic matter but which cannot be cultured and, secondly, autochthonous “epibacteria” which are amenable to culture, are larger than “planktobacteria” and utilize particulate organic matter. It will be difficult, however, to confirm this speculation and to establish unequivocally the presence of two distinct populations because so little is known of the physiology or structure of marine bacteria. Such information can only be obtained by studying large concentrations of single species populations, but the only way of obtaining enough material for experimentation is to culture the natural populations of bacteria. If “planktobacteria” really are different from other bacteria and cannot be cultured, then it will be almost impossible to do the phylogenetic studies which are necessary to prove that they are different. In the meantime, it is probably premature to consider these small bacteria as a distinct group since such a classification would be based on size and inability to grow in culture; these are poor taxonomic criteria, bacteria being notoriously pleomorphic and in the past, most fastidious bacteria have been cultured when the right growth conditions were supplied. There are reports of marine bacteria which can be cultured in the laboratory on natural organic carbon concentrations. Carlucci & Shimp (1974) reported one such isolate but Sieburth (1979) suggests that this was a saprophytic bacterium (one of his “epibacteria”) which had become adapted for growth on low organic concentrations and was not a true “planktobacterium”. Clearly, further study is required on the basic physiology of marine bacteria. There is some controversy as to whether marine bacteria are predominantly found attached to organic particles and detritus (Darnell, 1967; Odum & de la Cruz, 1967) or if they are predominantly free-living. In part, the problem was related to the difficulty in distinguishing bacteria from other particles but, even with the development of fluorochrome stains and scanning electron microscopy, there are still conflicting reports in the literature. Wiebe & Pomeroy (1972) and Taga & Matsuda (1974) found relatively few attached bacteria in neritic and pelagic waters and concluded that bacteria do not coat particles in the sea. Jannasch (1973) assumed that bacteria do colonize particles and interpreted the absence of bacteria on “marine snow” to mean that the particles were of very recent origin and had yet to be colonized; however, it could also be argued that this observation confirms the hypothesis that deep-sea bacteria do not attach to particles. Using scanning electron microscopy, Paerl (1975) found considerable attachment to particles in pelagic and nearshore waters and that the incidence of attachment was highest in near-surface waters with a supply of fresh organic matter from phytoplankton production. Goulder (1976) found the majority of bacteria in a turbid estuary to be attached. Jannasch & Pritchard (1972), in an experimental study with bacterial cultures, found that the density of added inorganic particles, as well as the concentration of dissolved organic matter, influenced the degree of attachment. There is, therefore, evidence both for and against the attachment of bacteria to particles. There is some attraction in Sieburth’s (1979) proposal that the small bacteria observed by epifluorescence microscopy are never attached to particles and utilize DOC, and that

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the larger bacteria, capable of culture, are those which do attach to particles and which utilize POC. Whether attached bacteria are ever capable of living free and whether the mode of nutrition affects the gross morphology of the bacteria is open to speculation and requires further study. CHEMICAL ESTIMATES Other methods of quantifying bacterial biomass are based on chemical analysis of cell constituents. One of the commonest indicators of biomass is adenosine triphosphate (ATP) which has been used extensively to estimate microbial biomass in marine surveys (Holm-Hansen & Booth, 1966; Holm-Hansen, 1969; Hobbie et al., 1972; Manuels & Postma, 1974; Devol, Packard & Holm-Hansen, 1976; Hodson, Holm-Hansen & Azam, 1976; Karl, Morse & Sturges, 1976; Karl & Holm-Hansen, 1978; Williams, Carlucci & Olson, 1980). As a means of estimating bacterial biomass, the technique has, however, severe limitations. It is usual to convert from ATP concentration to bacterial biomass using the carbon: ATP ratio determined by Hamilton & Holm-Hansen (1967) for bacterial cultures, which varied from 0.3–1.1% of the cell carbon depending on the physiological state of the cells. The size of the cell is also important and, as many marine bacteria are extremely small, it is necessary to correct for cell size in converting from ATP to bacterial biomass (Sieburth, 1979). The problem is further complicated because in the presence of phytoplankton cells, which are always found in the euphotic zone, it is impossible to distinguish bacterial from phytoplankton ATP. In a careful, statistical treatment of ATP levels and phytoplankton counts from a freshwater lake, Jassby (1975) showed that bacterial carbon could be estimated approximately only if the bacterial biomass was much greater than the phytoplankton biomass and only if the phytoplankton biomass was estimated with great precision. Therefore, ATP measurements should be confined to estimating bacterial biomass in deep water, outside the euphotic zone; unfortunately this restriction on the ATP method is all too often ignored by aquatic microbiologists. Clearly, a cellular constituent which is unique to bacteria would be the ideal means of estimating bacterial biomass in the sea; one candidate is muramic acid which has been found only in the cell walls of prokaryotic organisms, i.e. bacteria and blue-green algae (cyanobacteria). The method involves analysis of D-lactate released by acid hydrolysis of muramic acid and has been used successfully in marine sediments (Moriarty, 1975, 1977; King & White, 1977; Fazio, Mayberry & White, 1979); the concentration of bacteria in the water column, however, appears to be too low for the present sensitivity of the technique. Other indicators of bacterial biomass that have been proposed are lipid phosphate (King, White & Taylor, 1977; White et al., 1977, 1979) and certain branched chain fatty acids are valid markers for bacteria in marine sediments. Short chain C13–C21 iso and anteiso branch fatty acids are present in many bacteria (Kates, 1964; Kaneda, 1967, 1973; Parker, Van Boaler & Maurer, 1967; Tornabene, Gelpi & Oro, 1967; Tyrrell, 1968; Lechevalier, 1977) and the same acids are found in many recent sediments (Leo & Parker, 1966; Cooper & Blumer, 1968; Cranwell, 1973, 1974; Ishiwatari & Hanya, 1973; Johnson & Calder, 1973; Schultz & Quinn, 1973; Gaskell, Morris, Eglinton & Calvert,

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1975; Morris & Calvert, 1977). These branched acids have been found as major lipid components only in prokaryotic cells and their presence in sediments has been assumed to be due to bacteria (Leo & Parker, 1966; Cooper & Blumer, 1968). Internally branched fatty acids (i.e. 10 methyl C16 and C18) have been found in recent sediments by Cranwell (1973, 1974) and are thought to be the result of bacterial activity; these acids are, however, not exclusive to bacteria (Perry, Volkman, Johns & Bavor, 1979) and are not useful bacterial markers. Cyclopropane fatty acids (C17 and C19) have been found in certain gram-negative bacteria (Kates, 1964; Lennarz, 1970; Shaw, 1974; Lechevalier, 1977) and their presence in sediments is assumed to be the result of bacterial activity (Cranwell, 1974, 1976a,b, 1979; Order, 1976); the position of the cyclopropyl ring is, however, important as some isomers are also found in plants (Yaro, Nichols, Morris & James, 1972). Branched unsaturated acids (iso 15:1 and iso 17:1), first reported in a diatomaceous ooze by Boon, De Leeuw & Schenck (1975), were subsequently found in the sulphatereducing bacterium, Desulphovibrio sp. (Boon, De Leeuw, Hoek & Vosjan, 1977) and were proposed as a specific marker for Desulphovibrio. These acids have been found in a number of sediments (Matsuda & Koyama, 1977; Volkman & Johns, 1977; Perry et al., 1979) and, although they appear to be good markers of bacterial activity in sediments, their presence cannot be assumed to be due to a single bacterial species. Certain odd-chain mono-unsaturated fatty acids (e.g. 15:1 and 17:1) occur in sediments (Parker, 1967; Gaskell et al., 1975; Brooks et al., 1976; Matsuda & Koyama, 1977; Perry et al., 1979) and in some species of marine bacteria (Kunimoto, 1975; Kunimoto, Zarra & Igarashi, 1975) and have been suggested to indicate bacterial input of lipids into marine sediments. The geometry and position of double bonds in sedimentary fatty acids have been a valuable method of determining bacterial contributions (Matsuda & Koyama, 1977; Volkman & Johns, 1977). For example, cis-vaccenic acid (18:1 ω 7), reported in a number of sediments (Van Vleet & Quinn, 1976; Matsuda & Koyama, 1977; Volkman & Johns, 1977) is synthesized by the anaerobic, chain elongation pathway found predominantly in bacteria (Erwin, 1973) but the oxygen-dependent desaturases common to most organisms result in oleic acid (18:1 ω 9). Although cis-vaccenic acid is found in many organisms together with oleic acid, it appears to be present in considerably higher concentrations in bacteria and so is a useful marker for bacteria in sediments (Matsuda & Koyama, 1977; Volkman & Johns, 1977; Perry et al., 1979). Known biosynthetic pathways for monounsaturated fatty acids lead to cis isomers (Erwin, 1973) but there are a number of reports of trans isomers in marine sediments (Van Vleet & Quinn, 1976; 1979; Boon, De Leeuw & Burlingame, 1978). Perry et al. (1979) isolated trans monounsaturated acids from a tropical mangrove sediment and speculated that the trans fatty acids arose from bacterial activity; Volkman et al. (1980) implicated bacteria as the source of these compounds in an intertidal sediment. The occurrence of hydroxy-acids in sediments (Eglinton, Hunneman & ZadehDouraghi, 1968; Boon et al., 1977, 1978; Cardoso, Eglinton & Holloway, 1977; Volkman et al., 1980) has led to a number of suggestions as to possible sources including in situ oxidation of monocarboxylic acids (Eglinton et al., 1968) or a direct microbial input (Cardoso et al., 1977); there are, however, insufficient data at present for any

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definite conclusions (Volkman et al., 1980). The situation is also unclear as to the source of dicarboxylic acids which occur in sediments (Eglinton et al., 1968; Ishiwatari & Hanya, 1973, 1975; Johns & Order, 1975, Cranwell, 1978; Volkman et al., 1980) and the data indicate both in situ formation and a higher plant origin. Bobbie & White (1980) attempted to manipulate an estuarine detrital microbial community with antibiotics and altered culture conditions to stimulate fungal or bacterial growth and followed the changes in fatty acid methyl esters. Fungus enriched detritus showed enrichment of C18 dienoic, C18 and C20 polyenoic esters; an enrichment of the bacterial population resulted in a significantly larger proportion of anteiso and isobranched C15 fatty acid esters, C17 cyclopropane fatty esters and the cis-vaccenic isomer of the C18 monoenoic fatty acid esters. Although this approach seems to be useful in identifying the potential source of some of the compounds found in marine sediments, it may be that perturbation of a microbial community results in the production and accumulation of compounds which would normally be metabolized by another component of the community which is inhibited as a result of the perturbation. Stimulation of growth may also affect the production of potential marker compounds e.g. Bobbie & White (1980) comment that slowly growing or stressed bacteria accumulate cyclopropane fatty acids which are not found in rapidly growing populations. The basic limitation of all these marker compounds is that their presence in a bacterial species has to be shown by culturing the bacteria because it is impossible to remove bacterial cells from the natural environment for direct chemical determination; culturing a bacterial isolate is a poor way of categorizing the intermediates or end-products of metabolism, which may be greatly influenced by culture conditions (Tornabene, Bennett & Oro, 1967; Oliver & Colwell, 1973a,b). In spite of the obvious importance of bacteria to the process of diagenesis in marine sediments, little is known of the chemical composition or metabolism of natural populations of marine bacteria and there is a long way to go before we have a proper understanding of the rôle of marine bacteria in chemical diagenesis.

BACTERIAL PRODUCTION In addition to the absence of a general method for the determination of bacterial biomass, there is no single method which gives a reliable estimate of bacterial production. Studies of phytoplankton production received a considerable fillip when 14C was introduced by Steemann Nielsen (1951), because it is such a simple technique to use. Unfortunately, 14C-tracers cannot be used as easily in bacterial production studies because, unlike photosynthetic organisms which use only one substrate (CO2), bacteria can use many organic compounds. Since neither the composition of the utilizable organic matter nor its concentration are known, it is impossible to supply the same labelled compounds that the bacteria utilize in situ; 14C-organic compounds cannot do for bacterial studies what 14CO2 did for primary production. 14C

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It has become almost standard practice to use small molecular weight solutes, such as acetate, glucose or amino acids, labelled with 14C, to give an estimation of bacterial activity. There are two basic approaches; either the natural concentration of a chosen substrate is measured and the labelled compound is added as a true tracer (Williams & Askew, 1968; Williams, 1970; Williams & Gray, 1970; Andrews & Williams, 1971; Banoub & Williams, 1972; Williams, Berman & Holm-Hansen, 1976; Williams & Yentsch, 1976; Dawson & Gocke, 1978; Billen, Joiris, Wijnant & Gillain, 1980) or a kinetic approach is adopted which involves measuring the rate of uptake of substrate at several concentrations. This latter method gives a measure of heterotrophic potential but not of total activity; it was first proposed by Parsons & Strickland (1962) and modified by Wright & Hobbie (1965) for lake water and subsequently applied extensively in marine studies (Vaccaro & Jannasch, 1967; Hamilton & Preslan, 1970; Takahashi & Ichimura, 1971; Hobbie et al., 1972; Seki, Nakai & Otobe, 1972; Crawford, Hobbie & Webb, 1974; Seki, Yamaguchi & Ichimura, 1975; Sibert & Brown, 1975; Carney & Colwell, 1976; Gillespie, Morita & Jones, 1976; Albright, 1977; Gocke, 1977; Hoppe, 1978; Joint, 1978; Ramsey, 1978; Barvenik & Malloy, 1979). Both methods have advantages and disadvantages. The tracer method is superficially the preferred method since it gives an actual rate of incorporation and turnover of a specific substrate by the bacteria. The analytical problems involved in determining natural substrate concentration in sea water are, however, often formidable and analytical techniques are often not even available. Concentrations of readily utilizable organic solutes in the sea are generally <20µg·1−1 and often significantly less; e.g. Andrews & Williams (1971) reported glucose concentrations in the English Channel of 0.4–5.7 µg·1– 1 and total amino-acid concentrations of 20–80 µg·1−1 with individual amino-acid concentrations ranging from 0.4–8.1 µg·1−1. Such low concentrations obviously require very careful analysis and precautions must be taken against contamination of the samples. The analytical procedures for specific compounds tend to be laborious so that only a limited number of determinations can be achieved. Also, there are no convenient analytical procedures for many of the organic compounds which microbiologists assume to be preferentially utilized by bacteria, e.g. acetate, lactate, malate, succinate etc. More seriously, many interesting compounds are not available labelled at high specific activity; for example, glucose is available at about 80% 14C but malate and citrate are less than 20% 14C and glycollate is less than 10% 14C. It may be impossible to add sufficient 14C to get a reasonable count without adding so much 14C substrate that the natural substrate concentration is substantially increased and the added substrate would no longer be a tracer. The combination of difficult analytical technique and substrates available only at low specific activity, makes the tracer procedure less useful than one would hope. The second, kinetic, procedure was developed to allow an estimation of bacterial production without knowing anything about the natural substrate concentration. It involves adding a range of 14C substrate concentrations and measuring the assimilation of 14C into bacterial biomass and the production of 14CO22 by bacterial respiration. It is assumed that the uptake of substrate obeys Michaelis kinetics; a single reciprocal plot of substrate-dependent uptake rate against substrate concentration should give a straight line, the slope of which is equal to the reciprocal of maximum uptake velocity, Vm (Wright & Hobbie, 1965). The intercept of the line with the ordinate occurs at the natural

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substrate concentration and so gives a measure of turnover time of that concentration of substrate. One other parameter that can be obtained, the intercept on the absicca, Kt+Sn, is the sum of the transport constant (Kt) and the natural substrate concentrations (Sn). Allen (1969) used the technique to obtain an estimate of Sn in freshwater by subtracting an estimate of Kt, obtained from bacterial culture, from the derived value of Kt+Sn. This calculation is, however, of doubtful relevance because of the difficulty in extrapolating from laboratory derived values of Kt. Estimates of transport constant (Kt) derived for cultures of marine bacteria are very variable; Vaccaro & Jannasch (1966) reported values ranging from 3 to 106 µg glucose·1−1. It would obviously be foodhardy to attempt to estimate Sn using a value of Kt which was so imprecisely known. It is well known that pure culture of a single species growing on a single substrate exhibits Michaelis kinetics, but heterogeneous natural populations, utilizing several substrates, might not be expected to show the same kinetic properties. Williams (1973a) developed a simple mathematical model to test whether heterogenous populations deviate from Michaelis kinetics; he found deviations, especially at low substrate concentrations and the more diverse the population, the less successful was the kinetic approach. In practice, however, deviations from Michaelis kinetics are likely to be less than the inherent variability in the method and a heterogenous population can usually be assumed to behave like a single species. TURNOVER RATES OF ORGANIC MATTER BY BACTERIA Both the techniques using 14C solutes can give an estimate of the turnover of a variety of organic compounds in the sea. The range of reported values of turnover times is given in Table II. Generally, turnover is more rapid in estuaries and coastal waters than in the open sea and turnover times are significantly shorter in surface waters than at depth. A different approach to the estimation of organic turnover was taken by Ogura (1975) who measured temporal changes in DOC concentration in incubated samples of coastal water. He found an initial rapid decrease in concentration lasting between 1 and 5 days and a subsequent slower rate of decrease which was linear for 40 days. The initial decrease was assumed to be due to bacterial utilization of labile organic matter and the slower rate was due to the breakdown of more refractory organic compounds. This approach is, however, only likely to be useful in polluted waters which receive considerable amounts of allochthonous DOC; rapid changes in DOC concentration do not occur in most marine waters where the situation approximates to a steady state condition with production and utilization balancing to give very small changes in DOC concentration.

TABLE II Estimates of the turnover time of dissolved organic compounds obtained by the tracer and the kinetic approach

Depth Turnover

Oceanography and marine biology

Region

Substrate

Method

English Channel

14C glucose

Tracer

English Channel

14C glucose

14C amino acids

Pacific

Mediterranean

14C glucose

14C glucose

14C amino acids

Tracer

Tracer

Kinetic

Tracer

Tracer

(m)

101

(days)

Surface 6–66 10

1–37

10

04–>100

50

1.2–>100

10

3.5–>100

50

7–>100

Reference Williams & Askew (1968)

Andrews & Williams (1971)

Surface 11–17

Takahashi & Ichimura

50

44–149

(1971)

100

29–70

500

12–2041

1000

19–179

20

1.7–2.1

90

5–8.5

360

>100

1800

>100

20

3.2–3.9

90

5.2–8.2

360

>100

1800

>100

Banoub& Williams (1972)

14C glucose

Kinetic

50

250

14C aspartate

Kinetic

50

417

U.S.A., Estuary

14C glucose

Kinetic

Surface 0.01–0.93 Crawford et al. (1974)

Canada, Fjord

14C glucose

Kinetic

Surface 0.2–1.75

Sibert& Brown (1975)

California,

14C amino acids

Tracer

25

1.12–14

Williams et al. (1976)

100

7.8–>100

N. Pacific

Coastal Baltic

14C amino acid

Tracer

Surface 0.6–4.2 100

North Sea

14C alanine

Tracer

Seki et al. (1972)

Dawson & Gocke (1978)

2.7–13.6

Surface 0.91–97

14C glucose

Surface 0.95–18.9

14C acetate

Surface 0.34–69

14C lactate

Surface 2.1–104

Billen et al. (1980)

It is assumed in studies of organic turnover that uptake of labelled substrate is entirely by

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576

heterotrophic microbes and that phytoplankton are not capable of utilizing dissolved organic compounds at the concentrations present in the sea. Autoradiographic studies have generally shown no uptake of dissolved organics by phytoplankton (Munro & Brock, 1968; Horner & Alexander, 1972; Paerl & Goldman, 1972; Paerl, 1974) but there have also been reports of low rates of uptake by autotrophic organisms (Hoppe, 1976, 1977; Pollingher & Berman, 1976). Hoppe (1978) concluded, however, that heterotrophic activity of algae was very low and would have a negligible effect on measurements of bacterial production, but clearly, individual algal cells could benefit considerably from the assimilation of useful organic compounds which they did not have to synthesize. Other evidence for the importance of bacteria in heterotrophic processes in the sea comes from size fractionation. Williams (1970) found that 80% of added substrate was incorporated by small organisms which passed through 8 µm mean pore size filter and 49% was in organisms passing through 1.2 µm pore size. Derenbach & Williams (1974) found 80–97% of the heterotrophic production passed through 3 µm mean pore size filters and Larsson & Hagström (1979) found 95% passed through a 3 µm NucleporeR filter. This association of heterotrophic activity with the smallest size fractions suggests that bacteria, rather than protozoans or algae, are involved in the turnover of organic matter in the sea. MINERALIZATION In terrestrial ecosystems, the amount of primary production consumed by herbivores is much less than is generally assumed in marine environments; Gray & Williams (1971) estimated that only 6–40% of the autochthonous organic matter is eaten, the rest reaching the soil where it is available to decomposing microbes. Terrestrial microbes are, therefore, traditionally considered as re-mineralizers of organic matter. This rôle has also been ascribed to marine bacteria. Data obtained from 14C organic studies in the last decade cast doubts, however, on this assumption. Williams (1970) first pointed out that, in contrast to data obtained using bacterial cultures, a relatively small proportion of organic matter taken up by bacteria in the sea is respired and the major part of the carbon is assimilated into cellular material. That is, bacteria appear to be very efficient utilizers of organic substrates, which results in high growth yields, but are poor mineralizers. There is now considerable evidence in the literature to support Williams’ initial observation (Table III); generally, the percentage of glucose taken up which is respired is about 30%, giving a growth yield of about 70%. The growth yield from individual amino acids varies, with valine and leucine giving the highest growth yields (about 90%) and alanine, glutamate, asparate, serine, glycine, and arginine giving lower growth yields of 60–80% (Williams, Berman & Holm-Hansen, 1976). The data presented by Billen et al. (1980) are at variance with the other reports since they found growth yields of only 20–50% on glucose and 16–50% on a mixture of three amino acids. Using a different approach, Newell, Lucas & Linley (1981) have attempted to quantify the conversion of organic matter into bacterial biomass in culture by measuring the decrease in DOC and POC concentrations and concomitant increase in bacterial biomass; they estimated a conversion efficiency for carbon of only 10%, which is much lower than

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any estimate using 14C. Is it possible that the short incubation times employed in the 14C experiments are resulting in low estimates of bacterial respiration? How relevant are the laboratory experiments of the type done by Newell et al. (1981) to the situation in the deep-sea? Further experimentation, both in the laboratory and in situ, is required to clarify the question of growth efficiency of marine bacteria. Why should marine bacteria have such apparently high growth yields? Williams (1970, 1973b) suggested that the requirement for biosynthesis is reduced because bacteria have a range of organic compounds available which can be directly incorporated into macromolecules and biomass; in these circumstances, less energy is required for biosynthesis resulting in reduced catabolism and increased growth yields. High growth yields have implications for the traditionally conceived rôle

TABLE III Respiration of added substrates as a percentage of gross uptake

Region

Substrate

% respiration

Reference

English Channel

Glucose

24–27

Williams (1970)

Amino-acid mixtures

34

Glucose

33–46

Amino-acid mixtures

13–22

Glucose

28–49

Amino-acid mixtures

23–30

Glucose

8–17

Leucine

3–37

Alanine

34–47

Glutamate

42–57

Arginine

22–47

Valine

10–29

U.S.A.. Estuary

Glutamate

30–43

Carney& Colwell (1976)

North Sea

Glucose

28.9

Gocke (1976)

Acetate

34.8

Lactate

37.6

Malate

59.2

Amino-acid mixture

25.3

Amino-acid mixture

25.5

Glucose

20

Mediterranean

N.E. Atlantic

U.S.A., Estuary

Bahamas

Crawford et al. (1974)

Williams & Yentsch (1976)

Interlinking of physical

California, Coastal

Baltic

North Sea

Amino-acid mixture

15

Leucine

2–4

Alanine

31–35

Glutamate

20–30

Amino-acid mixture

19–31

Glucose

13–24

Glucose

50–80

Acetate

40–80

Lactate

40–85

Amino acids

50–84

578

Williams et al. (1976)

Dawson & Gocke (1978)

Billen et al. (1980)

of bacteria in recycling inorganic nutrients for subsequent utilization by phytoplankton. Terrestrial ecologists are well aware that bacteria can have a significant demand for inorganic nutrients and may be responsible for the removal of soluble inorganic nutrients; in the terminology of the soil microbiologist, such nutrients are immobilized. Decomposition of organic matter results in bacterial utilization of inorganic nutrients and incorporation into biomass; the C:N:P ratio of the organic matter influences the bacterial demand for inorganic N and P (Alexander, 1961) and hence the degree of mineralization or immobilization. Immobilization of nutrients by marine bacteria has rarely been considered (Thayer, 1974) but clearly it may be a significant process, especially if the high growth yields for carbon also apply to nitrogen. If the C:N and C:P ratios of organic matter are greater than those of the bacteria, then it will be necessary to supplement that organic matter with inorganic nutrients before there can be complete utilization. So under these circumstances, bacteria may compete with phytoplankton for available nutrients. Competition for phosphate has been reported by Rhee (1972), Faust & Correll (1976), and Friebele, Correll & Faust (1978) and competition for nitrogen and phosphorus by Thayer (1974). Some experiments have been done to measure nitrogen regeneration by marine bacteria. Hollibaugh (1978) supplied individual amino acids and measured the release of ammonia by an enrichment culture of marine bacteria; he found that the nitrogen regeneration ratio was about 0–8, that is, growth efficiencies in terms of nitrogen of only 20%. His estimates for carbon growth yield were also low (about 30%) which contrasts with the majority of reports of bacterial substrate utilization with their high growth yields. Hollibaugh (1978) used quite high concentrations of amino acid so that the ammonia produced could be chemically measurable and this may have influenced the results. Subsequent work by Hollibaugh, Carruthers, Fuhrman & Azam (1980) in enclosed water columns, gave high rates of bacterial nitrogen regeneration from natural organic matter. The bacterial production was closely coupled with, and appeared to be limited by, primary production. These experimental water columns were periodically supplied with inorganic nutrients and it is difficult to extrapolate from this situation to the nutrientlimited ocean. So, the data of Hollibaugh (1978) and Hollibaugh et al. (1980) do suggest

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that bacteria are important in re-mineralization of nitrogen but the data of others (Table III) suggests a low mineralization rate of carbon compounds. Clearly, further studies on nitrogen and carbon turnover, in both nutrient-rich and nutrient-depleted sea water, are required to resolve this discrepancy. COUPLING BACTERIAL AND PRIMARY PRODUCTION Within the euphotic zone, production of DOC by phytoplankton may be a major source of readily metabolizable substrate for bacteria and the interactions occurring between bacteria and phytoplankton have been the subject of several studies. Initially, experiments involved supplying bacterial populations with known products of phytoplankton excretion, such as glycollate (Wright, 1970; Tanaka, Nakanishi & Kadota, 1974) but in recent years, several studies have measure simultaneous phytoplankton production of DOC and DOC uptake by bacteria. Experimental approaches have been quite varied; Derenbach & Williams (1974) and Larsson & Hagström (1979) used differential filtration through membrane filters of different pore size to separate bacteria from phytoplankton. Larsson & Hagström (1979) found, after 4 h incubation, that 65% of the labelled carbonate supplied was found in phytoplankton cells, 27% in bacterial cells, and 8% was DOC; the very rapid release of organic matter by phytoplankton cells was, therefore, simultaneously taken up by bacteria. Berman (1975) and Iturriaga & Hoppe (1977) attempted to differentiate heterotrophic and autotrophic production by supplying antibiotics to inhibit bacterial activity and found that DOC release was between 1 and 10% higher in the presence of antibiotics; they assumed this increase was due to the suppression of bacterial utilization. These studies are, however, less than clear-cut because of doubts over the efficiency of antibiotics in immobilizing marine bacteria; it is not possible to estimate the flux of carbon from phytoplankton to bacteria using this method. Another approach was adopted by Wiebe & Smith (1977) and by Schleyer (1980) which involved supplying 14C-labelled algal extract as tracers. Working in an Australian estuary, Wiebe & Smith (1977) found very close coupling between phytoplankton exudation and bacterial uptake and suggested that the rapid bacterial utilization would make it impossible for phytoplankton exudation ever to be directly responsible for the DOC found in marine waters. Interestingly, most of the uptake of 14C-DOC was by particles of 100–124 µm in diameter, suggesting either that heterotrophs responsible for uptake in this estuary were not bacteria, which is unlikely, or that bacteria attached to particles were responsible. Schleyer (1980) adopting the kinetic approach of Wright & Hobbie (1965), used labelled algal extract instead of a defined substrate. There were, however, several problems which made interpretation difficult; he found that the procedure over-estimated utilization of natural DOC because an unknown proportion of organic compounds are resistant to microbial degradation. Also, the production estimates made by this method were much greater than the increase in biomass measured by epifluorescence microscopy. Schleyer obtained growth yields on labelled algal organics which were much higher than those obtained with glucose; only 1.5% of the algal DOC was respired by the bacteria, giving a growth yield of 98.5%; such a high growth yield seems improbable and, even if the assimilated organic compounds were utilized in

Interlinking of physical

580

anabolism as Williams (1970) suggested, it would be remarkable if the energy maintenance requirements of the bacteria required a respiration rate of only 1.5% of uptake. Clearly, the bacteria must have been utilizing other, unlabelled organic compounds to meet their maintenance requirements. Another method of determining bacterial utilization of phytoplankton DOC was used by Smith, Barber & Huntsman (1977) who measured the different rates of accumulation of DOC in the light and dark in a coastal upwelling off northwestern Africa; their carbon flux model suggested that bacteria remove 18% of the DOC excreted by phytoplankton per hour, but this represented only 1.7% of the total autotrophic production of organic matter. Clearly, this approach is open to criticism if, as Wiebe & Smith (1977) suggested, there is close coupling between phytoplankton DOC excretion and bacterial utilization, because DOC accumulation may be only a small fraction of the total flux between phytoplankton and bacteria. Smith et al. (1977) specifically discounted any rapid and complete removal of phytoplankton exudate by bacteria. A similar approach of relating diurnal changes in DOC with heterotrophic activity was used by Sieburth, Johnson, Burney & Lavoie (1977); they found that, in regions of intense microbial activity, there was a difference of 13% in DOC concentration and 32% in total carbohydrate between day and night samples. This result implies that bacterial utilization of algal DOC must be greater than Smith et al. (1977) found but much less than the rates reported by Wiebe & Smith (1977). Clearly, there are two questions which future studies must answer; how important is phytoplankton exudation of DOC to the total pool of DOC available for bacterial utilization and, are the reported differences in the coupling of phytoplankton exudation and bacterial uptake due to experimental artefacts or are there regional differences in the proportion of algal excretion which is directly utilized by bacteria Recently, Fuhrman, Ammerman & Azam (1980) studied bacterial production in a coastal euphotic zone and found that bacterial growth rate was related more to the standing stock of phytoplankton than to their production. It may be that bacteria were utilizing organic compounds produced by zooplankton predation rather than using DOC excreted during photosynthesis. If this result is found in other regions, then zooplankton grazing may play a dominant rôle in supplying organic matter for bacterial growth.

TROPHIC RÔLE OF MARINE BACTERIA The classic view of the marine food web is of three or four trophic levels from phytoplankton to fish (Ryther, 1969) but it is only recently that the place of marine bacteria in this scheme has been considered. A considerable stimulus was provided by the work of Russian workers who claimed that bacterial production was often greater than primary production. Sorokin (1971a, b, 1973, 1977) reported that the biomass and production of bacteria in the tropical Pacific considerably exceeded primary production. Apart from one report of bacterial production being less than primary production in a winter diatom bloom under ice in the Sea of Japan (Sorokin & Konovalova, 1973), Sorokin’s results consistently suggested that bacterial, heterotrophic production was greater than phytoplankton, autotrophic production. The obvious question is, where does

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the organic matter come from to fuel this bacterial production? Sorokin (1971b) suggested the source was advection of deep-water from the more productive regions of the ocean into the area of the tropical Pacific that he was studying. Banse (1974) criticized this hypothesis because it could not be reconciled with the observed geographical distribution of DOC. Banse also reasoned that bacterial production could not be greater than 10% of phytoplankton production and that Sorokin’s estimates of bacterial production must be too high by at least an order of magnitude. The procedure used by the Russian workers to estimate heterotrophic production is open to criticism. The technique, first proposed by Romanenko (1964), is based on the fixation of 14CO2 by heterotrophic processes and assumes a constant relationship between dark 14CO2 fixation and total heterotrophic assimilation. Such dark CO2 fixation, however, is not strictly analagous to autotrophic carbon reduction because there is evidence (Overbeck, 1972) that the CO2 is fixed in anaplerotic reactions; these reactions are involved in replenishing intermediates of the TCA cycle which are diverted into biosynthesis (Kornberg, 1966) and this CO2 fixation does not represent de novo synthesis of biomass. The amount of carbon fixed in anaplerotic reactions varies considerably with substrate utilization and it would be improbable if there were a constant relationship between dark CO2 fixation and bacterial production. Other problems, such as the certain dark CO2 fixation by algae and the inconstancy of the ratio of dark CO2 fixation to total CO2 fixation have been discussed by Overbeck & Daley (1973) who cautioned against using the Romanenko technique. Yet, it would be premature to dismiss out of hand the Russian workers’ results on the basis of a doubtful technique. Other workers have also presented evidence that bacterial production is much greater than has previously been assumed. Joiris (1977) reported that heterotrophic consumption of organic matter in the southern North Sea was ten times greater than primary production. In considering all the possible sources of error in his methods, Joiris concluded that his estimate of heterotrophic respiration was unlikely to be an under-estimate but he had less confidence in the measurements of primary production, which were based on changes in dissolved oxygen concentration. Sieburth et al. (1976) suggested that 30–40% of the plankton biomass in the north Atlantic was composed of bacteria and that their estimates of bacterial production agreed well with Sorokin (1971b). This bacterial production could not be sustained by the estimates of primary production which are obtained using the standard 14C technique (Sieburth, 1977). It is generally accepted that the 14C method may be under-estimating phytoplankton production (see review by Peterson, 1980) but it is hardly credible that a technique which has been so widely used should under-estimate primary production by an order of magnitude as these estimates of bacterial production imply. Sieburth (1977) thought that the bacterial production estimates were more reliable because they were based on a number of different techniques, whereas primary production is measured almost exclusively by the 14C method. None of the methods used to measure bacterial production is, however, free from error and, perhaps, an error of two or three in both heterotrophic and autotrophic estimates may remove these apparent discrepancies between bacterial and phytoplankton production. It may also be that the above arguments are misleading because they relate bacterial production to primary production, implying that bacteria are competing directly with

Interlinking of physical

582

zooplankton for the organic matter produced by phytoplankton. This need not be the case and an alternative hypothesis which depends on the efficiency with which zooplankton utilize ingested phytoplankton carbon has already been discussed (Fig. 1, p. 70). Such a process, coupled with phytoplankton production which we have previously speculated may not be completely utilized by zooplankton because of temporal mismatch in biomass, could result in a large proportion of the production passing to the DOC- and POC-pools. The end result is that a significant proportion of the total primary production (although much less than 100%) could be utilized by bacteria, without depriving zooplankton of any food. GRAZING ON BACTERIA More abundant bacterial biomass and greater bacterial production inevitably raises the question of what happens to the bacteria and is there a component of the food web which can exert a significant grazing pressure on marine bacteria? Boyd (1976) stated that few copepods are capable of filtering particles <5 µm diameter and, in a detailed study of grazing by Acartia clausi, Nival & Nival (1976) found that the mean diameter of particles removed by young stages (copepodite I) was 3 µm which increased to 7 µm with adult animals; the minimum particle diameter which was retained by copepodite stage I was 2 µm. It is, therefore, unlikely that filter-feeding copepods could remove free living bacteria, especially if they were the small (<0.4 µm diameter) bacteria commonly reported in the sea. Clearly, bacteria attached to particles could be removed by copepods and Sorokin (1973) found grazing on bacterial clumps in the tropical Pacific. Rieper (1978) showed that bacteria could be a significant food source for harpacticoid copepods and King, Hollibaugh & Azam (1980) found that bacteria could form a substantial part of the diet of the larvacean, Oikopleura dioica, but this grazing had very little effect in controlling the numbers of bacteria present. There have been very few experiments on the flow of carbon through marine food webs which have quantified the trophic rôle of bacteria. Smith et al. (1979) followed 14Ccarbonate and 3H2O through a food web involving demersal (epibenthic) zooplankton and found that phytoplankton did not appear to be the main food source; they speculated that bacteria associated with particles represented a major source of food. Brown & Sibert (1977) also implicated bacteria and found that benthic harpacticoid copepods ingested bacterial carbon 8–10 times faster than algal carbon when fed enrichment cultures of bacteria and benthic algae. All too often the possible significance of bacteria as food is, however, ignored; in their experiments to determine the carbon budget of Calanus pacificus feeding on unialgal diet labelled with 14C, Copping & Lorenzen (1980) did not even mention bacteria, although considerable quantities of DOC were released from the labelled phytoplankton over the 48-h experimental period. Such significant supplies of DOC must surely have supported a large bacterial population but Copping & Lorenzen (1980) did not consider the possibility of bacteria as a food source for Calanus.

EFFECT OF STARVATION ON BACTERIAL ACTIVITY

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Most studies of bacteria are concerned with the rate at which new biomass is synthesized or organic matter is turned over by actively growing bacteria. Very few studies have considered the behaviour of bacteria which are not actively growing, yet there is increasing evidence from the low concentration of utilizable organic matter in the sea (Jørgensen, 1976) that bacteria in the open sea are in a state of near-starvation. There have been several studies of the response of bacterial cultures, especially of Klebsiella (Aerobacter) aerogenes, to starvation (Postgate & Hunter, 1962, 1963a, b, 1964); they showed that, if a culture is growth-limited by the exhaustion of a single substrate, then subsequent supply of that substrate results in death of the organism. That is, the presence in the recovery medium of a substrate which previously sustained growth, and would appear to be the ideal nutrient, in fact causes the death of the bacterium; this phenomenon is known as substrate-accelerated death and has been demonstrated with carbon-, phosphate- and ammonium-limited cultures in a number of bacteria including K.aerogenes, Escherichia coli, Streptococcus lactis, Azotobacter vinelandii, Arthobacter spp. and Mycobacterium fortuitum (Postgate & Hunter, 1962, 1963a, b, 1964; Dawes, 1976; Majtan & Drobrica, 1979; Calcott & Calvert, 1981). It is not known whether substrate-accelerated death occurs in the natural environment where populations presumably utilize a number of organic compounds and are unlikely to be growth-limited by a single substrate. The problem facing all bacteria when the supply of nutrients is exhausted is the same; all cellular processes such as osmotic and ionic regulation, turnover of structural macromolecules, etc. require energy and when an energy source is not available to maintain these processes the cell dies. Pirt (1965) made some estimates of maintenance energy requirements for Escherichia coli and Klebsiella aerogenes and found that between 0.07 and 0.09 µg glucose were required each hour by 1 g dry wt of cells. An extrapolation from these data to the open ocean where typical bacterial biomass might be about 0.1µg C·1−1, suggests that the equivalent of 0.5 µg glucose·1−1 would be required every day to meet the maintenance energy requirement of this population. Such an organic demand intuitively seems too high to be realistic and marine bacteria must have a much lower maintenance energy requirement than Escherichia coli or Klebsiella aerogenes. It is, however, difficult to establish experimentally the maintenance requirement of a natural population and calculations must be based on natural substrate concentrations; Williams (1975) quoted glucose concentrations ranging from undetectable to 20 µg·1−1 so the daily maintenance requirement extrapolated from K.aerogenes would quickly deplete the observed glucose concentrations, without resulting in any growth of the bacteria. Clearly, a number of substrates are available for the bacteria to utilize but such a high maintenance requirement could result in a rapid flux of dissolved organic compounds just to keep the observed population alive but not growing. The problem of deciding from a chemical measurement, if an organic compound is present in sufficient quantity to support a bacterial population may be further complicated by the apparent presence of bacteriostatic compounds in certain waters. Sieburth (1971b) reported the presence of a bacterial inhibitor in the surface film. Similar inhibition was found by Bell, Lang & Mitchell (1974) but they also found that excretion products of Skeletonema costatum could stimulate growth of some bacterial species in culture; they

Interlinking of physical

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suggested that a combination of stimulation and inhibition by extracellular products of phytoplankton could be a mechanism of controlling the dominant bacterial species. Such control mechanisms have not, however, been reported in the open ocean. Novitsky & Morita (1976, 1977, 1978a) reported an interesting effect of starvation on a psychrophilic marine vibrio, isolated from the Antarctic Convergence, which changed shape from a rod to a coccus and became much smaller (<0.4 µm diameter). The numbers of viable cells increased 8-fold during the first week of starvation and approximately 50% of this population remained viable during 2 months starvation; some cells were still viable after a year. Therefore, the survival mechanism of this marine bacterium appears to produce a large number of very small cells which have a very greatly reduced endogenous respiration rate (<1%). Similar reductions in respiration rate have been reported when soil bacteria are starved (Boylen & Ensign, 1970; Ensign, 1970). Novitsky & Morita (1978a) speculated that the increase in the number of cells may be a mechanism which ensures the survival of a bacterial species at times of starvation by the rapid proliferation of genetic material. Further work is, however, required to see how widespread such phenomena are in marine bacteria.

EFFECT OF BACTERIAL ACTIVITY ON THE INORGANIC CHEMICAL COMPOSITION OF THE SEA Although this review is primarily concerned with carbon turnover in the sea, bacteria may also have a significant effect on the inorganic constituents of sea water, in particular, the chemistry of nitrogen and sulphur compounds. Chemoautotrophic bacteria utilize reduced sulphur and nitrogen in the fixation of carbon dioxide and as a result, oxidized sulphur and nitrogen compounds are released into the marine environment; under certain conditions, bacteria utilize oxidized nitrogen and sulphur during heterotrophic growth on organic compounds. It is, therefore, pertinent to consider the alteration in the chemical composition of the sea as a result of bacterial activity. MARINE CHEMOAUTOTROPHIC BACTERIA Evidence has recently been presented that sulphur chemoautotrophs can contribute significant primary production in marine sediments: Kepkay, Cooke & Novitsky (1979), working in the sulphide-rich muds of Halifax Harbour, Nova Scotia, found a peak of CO2 fixation at a depth of 40 cm below the sediment surface, coincident with the maximum concentration of organic carbon in the sediment. Kepkay & Novitsky (1980) subsequently showed that the peak of CO2 fixation occurred at the transition zone between aerobic and anoxic mud and that there was an increase in sulphate concentration and a pH minimum at this depth. These data strongly suggested a substantial population of sulphur chemoautotrophic bacteria in the marine sediment; 210Pb data showed that the sediment had been accumulating at a constant rate of 2.5 mm·yr−1 implying that the observed organic carbon matter at 40 cm had been generated over the last 120 years. The measured chemoautotrophic CO2 fixation rate was much higher than that required to accumulate the observed organic carbon over this time and mass balance calculations

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indicated that 69% of the chemoautotrophically generated organic carbon had been mineralized. Kepkay & Novitsky (1980) speculated that, because of the long time scale involved in the production of the carbon peak and because of the undisturbed nature of the sediment, very little of the chemoautotrophically produced carbon could be utilized by higher trophic levels. In a subsequent paper, Novitsky & Kepkay (1981) found elevated heterotrophic production at the depth of the organic carbon peak (40 cm) which was fuelled by DOC release by the chemoautotrophs; 49% of the carbon fixed by the sulphur bacteria was released as soluble organic compounds. It is not known how common such examples of chemoautotrophic primary production are in marine sediments but they will probably be quite widespread; chemoautotrophic sulphur bacteria require a source of reduced, inorganic sulphur as well as molecular oxygen and conditions favourable to their development must be common in most marine sediments and in certain marine waters with anoxic deep layers. Sorokin (1970) described one such environment in the Black Sea where the fixation of CO2 and the simultaneous oxidation of H2S proved conclusively that chemoautotrophic sulphur bacteria were active. The Black Sea is a unique environment since anoxic water occurs in the euphotic zone and allows the development of photosynthetic sulphur bacteria (Dickman & Artuz, 1978) which are quite rare in the sea. Nitrification is the process of chemoautotrophic growth involving reduced nitrogen as electron donor and results in the autotrophic fixation of CO2 with the oxidation of ammonia to nitrite or nitrite to nitrate. In the past, the presence of high concentrations of nitrite in sea water was assumed to be the result of nitrifying bacteria (Brandhorst, 1959); early experiments with cultures of marine nitrifying bacteria found, however, very low rates of nitrite production which could not be ecologically significant (Carlucci & Strickland, 1968; Carlucci & McNally, 1969). Carlucci, Hartwig & Bowes (1970) concluded that nitrification might be significant in certain micro-environments with high ammonia concentrations but they also showed that nitrite could be released by phytoplankton in conditions of excess nitrate; observations of high nitrite concentration could, therefore, be the result of either or both these processes. These studies were based on extrapolations from laboratory experiments with cultures of marine nitrifying bacteria and, until recently, there was no sensitive method for directly measuring nitrification rates in the sea. There is, however, now a method of measuring chemoautotrophic CO2 fixation due to nitrifying bacteria which involves incubating water samples in the presence of an inhibitor of nitrification. The inhibitor “NServe”, 2-chloro-6-(trichloromethyl) pyridine, was developed to conserve ammonia fertilizers applied to agricultural land by specifically inhibiting nitrifying bacteria (Goring, 1962; Campbell & Aleem, 1965). The method relies on the differences in 14CO2 fixation between samples incubated in the presence and absence of “N-Serve” and gives a direct measurement of primary production by nitrifying bacteria. Measurements of nitrifying using this technique or the older method of measuring changes in concentration of ammonia, nitrite and nitrate, have so far been used mostly in coastal or estuarine sediments (Vanderborght & Billen, 1975; Grundmanis & Murray, 1977; Vanderborght, Wollast & Billen, 1977; Billen, 1978; Koike & Hattori, 1978b; Henriksen, 1980). The emphasis of such studies is usually on nitrogen diagenesis and the contribution of nitrifying bacteria to primary production in sediments is rarely considered.

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HETEROTROPHY WITH NITRATE AND SULPHATE AS TERMINAL ELECTRON ACCEPTORS Although the major part of the ocean has dissolved oxygen concentrations high enough for aerobic metabolism of bacteria, there are environments which have low dissolved oxygen concentrations where nitrate and sulphate may act as terminal electron acceptors of bacterial respiration. Nitrate and sulphate reduction are common in littoral sediments and in certain marine basins and upwelling regions. Fiadeiro & Strickland (1968) described the upwelling region off Peru where, at certain depths, the large input of organic matter by primary production resulted in high rates of respiration which reduced dissolved oxygen concentration (<0.2 ml O2·1−1) and produced a concomitant nitrite maximum. The contribution of nitrate-dependent heterotrophic production was, however, small compared with total heterotrophic activity and the amount of organic matter used in dissimilatory nitrate reduction was only 2–4% of the organic matter respired to produce the observed oxygen minimum. Fiadeiro & Strickland (1968) also commented that the concentration of organic matter available in these waters was significantly less than the concentrations which are found to fuel nitrate reduction in lakes and sediments where nitrate reduction is a significant process and they were surprised that the DOC concentration was high enough to support nitrate reduction. Carlucci & Schubert (1969), however, isolated a nitrate-reducing bacterium from these waters which was capable of reducing nitrate in sea water with no supplemental organic matter; therefore, bacteria were present which could reduce nitrate at the low organic concentrations found by Fiadeiro & Strickland (1968). Denitrification, that is the complete reduction of nitrate to a gaseous end-product (N2 or N2O) has been studied in upwelling regions of the eastern Pacific. Goering (1968) used 15N to demonstrate production of molecular nitrogen as a result of denitrification of nitrate but the rates he estimated were rather high, probably because of the extended incubation periods he had to employ to get measurable production of 15N2. In a more recent study, Dugdale et al. (1977) reported a spectacular example of denitrification in the Peru upwelling when nitrate was completely utilized and sulphate respiration occurred with the production of H2S; such examples of complete denitrification are rare in the open ocean but have been described in anoxic basins such as the Black Sea, Cariaco Trench, and fjords (Richards, 1965). Packard, Dugdale, Goering & Barber (1978) measured nitrate reductase activity in the Peru upwelling region and distinguished between the phytoplankton enzyme in the surface waters and that of bacteria in deeper water where the nitrite maximum occurred. They also calculated that denitrification produced 7 µg-at.N·1−1 of NO, N2O or N2 from an initial nitrate+nitrite concentration of 27.5 µg-at.N·1−1. Denitrification, therefore, occurred simultaneously with nitrate reduction and nitrite accumulation. Recently, Codispoti & Packard (1980) speculated that denitrification rates in the oxygen deficient waters of the eastern Pacific have increased since 1972 as a result of more of the products of primary production sinking out of the euphotic zone into the deeper, oxygen poor waters where denitrification occurs. These authors suggested that this may be one of the ecological changes which occurred in the 1970s as a result of the natural variation in upwelling produced by El Niño, the warm water intrusion into the

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upwelling region which decimated the anchovy fishery. In a subsequent paper Pak, Codispoti & Zaneveld (1980), however, suggested that the supply of organic matter to the denitrifying layers was not due to “raining down” of primary production but was due to horizontal, offshore transport of particulate matter. Denitrification has also been studied extensively in sediments; it occurs in anoxic sediments below the surface activity of aerobic heterotrophs and nitrifying bacteria (Goering & Pamatmat, 1971; Vanderborght & Billen, 1975; Heitzer & Ottow, 1976; Bender et al., 1977; Grundmanis & Murray, 1977; Vanderborght, Wollast & Billen, 1977; Billen, 1978; Koike & Hattori, 1978a, b, 1979; Søfrensen, 1978a, b; Oren & Blackburn, 1979; Seitzinger, Nixon, Pilson & Burke, 1980). The loss of fixed nitrogen as a result of benthic denitrification, fuelled by organic matter in sediments, may be quite significant and has been estimated to be 1.6×1013 g N·yr−1 by Codispoti & Richards (1976), 5×1013 g N·yr−1 by Bender et al. (1977) and 1.5×1013 g N·yr−1 by Koike & Hattori (1979). Considerable research has also been done on the strictly anaerobic heterotrophic processes involving the reduction of sulphate to sulphide, which occur extensively in marine sediments (Fenchel & Riedl, 1970; Nedwell & Floodgate, 1972; Engvall & Hallberg, 1973; Jørgensen & Fenchel, 1974; Ramm & Bella, 1974; Goldhaber & Kaplan, 1975; Jørgensen, 1977a, b, 1979; Berner, 1978; Sørensen, Jørgensen & Revsbech, 1979). Within marine sediments, sulphate reduction can account for more than 50% of the total mineralization of organic matter by bacteria (Jørgensen & Fenchel, 1974; Jørgensen, 1977a). Even although sulphate-reducing bacteria are strictly anaerobic, bacterial sulphate reduction has been reported in oxidized sediments which have localized anoxic regions and a large supply of organic matter (Jørgensen, 1977b). The reactions occurring in the diagenesis of sulphur as a result of bacterial activity are outside the scope of this review but Goldhaber & Kaplan (1980) give a recent review. METHANE PRODUCTION IN ANOXIC ENVIRONMENTS Methane production represents the final stages in organic matter mineralization in anoxic environments. Anaerobic fermentative reactions result in the production of simple organic molecules from more complex substrates, e.g. lactate is utilized by sulphatereducing bacteria and acetate is released; therefore, a variety of small molecular weight compounds are produced as a result of bacterial fermentation within the sediment and each compound may be the substrate for a group of bacteria or for a single species. When sulphate is depleted in deeper sediments below the layers of active sulphate reduction, methanogenic bacteria reduce CO2 to methane using the reducing power of only the very simplest organic compounds; indeed, cultures of methanogens are fastidious and none has been shown to produce methane with organic molecules more complex than acetate (Bryant, Campbell, Reddy & Crabill, 1977; McInerney, Bryant & Pfennig, 1979). An excellent review of methane cycling in marine and fresh water has recently been produced by Rudd & Taylor (1980). There is a clear link between sulphate-reducing bacteria and methanogens in marine sediments; the highest concentrations of methane are always found in the deepest layers of the sediment where sulphate concentrations are only about 10% sea-water

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concentration (Nissenbaum, Presley & Kaplan, 1972; Claypool & Kaplan, 1974; Martens & Berner, 1974, 1977; Martens, Berner & Rosenfeld, 1978). Neither sulphate, nor the product of sulphate reduction, sulphite are inhibitory to methanogens but the vertical distribution in sediment seems to be the result of the ability of sulphate-reducing bacteria to out-compete with methanogens for the available organic substrate; it is only under conditions of sulphate-limitation that production by methanogenic bacteria becomes significant. Organic substrate-limitation of methanogenic bacteria in sediments containing significant amounts of sulphate was shown by Winfrey & Zeikus (1977) who stimulated methanogenic bacterial production by adding acetate to the sediment. There are differences in methane cycling in freshwater sediments and marine sediments but these appear to be differences in the utilization of methane, rather than in methane production. In freshwater sediments, aerobic utilization of methane appears to be the dominant reaction but in marine sediments, anaerobic oxidation of methane by sulphur bacteria appears to control methane concentration (Reeburgh & Heggie, 1977); this presumably reflects the different sulphate concentration in freshwater sediments and marine sediments, with a higher concentration of terminal electron acceptor in marine sediments resulting in more heterotrophic utilization of organic matter. A combination of competition between methanogens and sulphate-reducing bacteria for organic substrate and utilization of methane by sulphate-reducing bacteria results in the depth distribution of methane found in sediments, with the highest concentrations occurring below the layers of active sulphate reduction. Methane production also occurs in the water column of anoxic marine basins and methane cycling in the Cariaco Trench has been studied by several workers. Reeburgh (1976) combined the fluxes of methane, CO2, organic carbon and sulphate to produce a budget of methane cycling; he found the upward flux of methane to be approximately equal to that of CO2, which is what one would expect if methane were produced by acetate fermentation. The consumption of methane by sulphate-reducing bacteria in the water column accounted for 85% of the methane flux. Recently, Romesser et al. (1979) isolated methanogenic bacteria from the Black Sea and Cariaco Trench and identified two new species, and a new genus, of methanogenic bacteria.

DISTRIBUTION OF BACTERIA IN THE SEA The horizontal and vertical distribution of bacteria in the sea is patchy. In the open ocean, numbers of bacteria are highest in the euphotic zone and decrease considerably below about 200–500 m (Packard, Healy & Richards, 1971; Sieburth, 1971a; Takahashi & Ichimura, 1971; Hobbie et al., 1972; Gundersen et al., 1972; Sieburth et al., 1977; Karl & Holm-Hansen, 1978; Williams, Carlucci & Olson, 1980). Generally, numbers are higher in neritic, coastal waters and estuaries (Seki, 1971; Erkenbrecher & Stevenson, 1975; Ferguson & Rublee, 1976; Palumbo & Ferguson, 1978; Ferguson & Palumbo, 1979) and considerable bacterial biomass is also present in shallow water sediments (ZoBell & Anderson, 1936; Bianchi, 1973; Boeyé, Wayenbergh & Aerts, 1975; Litchfield & Floodgate, 1975; Hanson & Gardner, 1978; Nair, Bharathi & Achuthankutty, 1978). Most bacterial activity occurs in or near the euphotic zone but there are different

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environments within the sea which pose special problems for bacteria. Although not quantitatively important in terms of total carbon turnover, some special environments will be considered because of the unique conditions found there; these are the oceanic surface film, the deep-sea water column, deep-sea sediments, hydrothermal vents, and the sediments of the upwelling region off Peru. BACTERIA IN THE SURFACE FILM It has long been recognized that the interface between the ocean and the atmosphere has some unique properties; even before there was extensive oil pollution, “slicks” were often observed on calm seas. It is, however, only comparatively recently that techniques have been developed which allow the surface microlayer to be separated from the rest of the water column (MacIntyre, 1974). Interest first developed in the chemistry of the surface film with its high concentration of naturally occurring non-polar organic compounds and petroleum hydrocarbons derived from both the atmosphere and the water column (Barbier, Tusseau, Marty & Saliot, 1981). The chemistry of this surface microlayer has been reviewed by Liss (1975). Detailed analysis of natural-product lipid constituents have been made by Garrett (1967), Kattner & Brockmann (1978), and Marty et al. (1979), who studied various lipid constituents removed by chloroform extraction. DOC has been reported to be concentrated in the surface microlayer up to 5 times, relative to subsurface water (Goering & Menzel, 1965; Williams, 1967) but the concentration is not uniform and even within slicks Williams, Van Vleet & Booth (1980) found the surface film to be very patchy. The first study of the bacterial composition of the sea surface microlayer was made by Sieburth (1965) who found numbers of culturable bacteria in the surface film which were two orders of magnitude greater than in the immediate subsurface water. Subsequently, Sieburth (1971a) studied the ability of bacteria isolated from the surface film to utilize lipid, protein, and carbohydrate; in the Pacific, surface film bacteria had consistently higher lipolytic and proteolytic activity than bacteria from subsurface water but this was not the case with surface film bacteria from the Atlantic and Caribbean which showed greater activity with carbohydrate. It is not clear why there should be this regional difference in the potential activity of surface film bacteria, unless the composition of the surface film was very different. Bacteria are not the only organisms found at high concentration in the surface film. Harvey & Burzell (1972) found significant concentrations of dinoflagellates and unidentified µ-flagellates but very low numbers of diatoms and ciliates; Crow, Ahearn & Cook (1975) reported high fungal and bacterial biomass in coastal surface films and Gallagher (1975) found that 37% of net plankton photosynthesis was concentrated in the surface film of a salt marsh community. Carty & Colwell (1975) reported that the numbers of viable bacteria in the surface film decreased in proportion to the distance from land and that enrichment in the surface film was between 1.3 and 6 times the concentration in subsurface water. Sieburth et al. (1976) calculated the probable cell densities in the surface film by assuming the thickness of the organic film to be 0.1 µm and that bacteria are concentrated in a zone 1.0 µm thick; they suggested that bacterial densities and organic concentrations

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in the surface film would be similar to those found in laboratory cultures of bacteria, i.e. organic matter concentration equivalent to 2.9 g·1−1 and bacterial biomass between 0.08– 5 mg C·1−1. Their estimates of the thickness of the surface microlayer were based on the data of Baier, Goupil, Perlmutter & King (1974) who found that the major constituents of the surface film were not the expected lipid, but glycoproteins and proteoglycans. The estimates of Sieburth et al. (1976) on bacterial population densities are difficult to confirm because of the problems of sampling the surface film without destroying its inherent structure. Sieburth et al. (1976) found a good correlation between numbers of bacteria and amoebae and assumed that amoebae were feeding on bacteria. Simultaneous high biomass of bacteria and amoebae, however, would be just as likely if they had a common mechanism of accumulation in the surface film; indeed, if amoebae were grazing bacteria one might expect a poor correlation since the amoebae would reduce the bacterial numbers and destroy any correlation. Are the bacterial numbers high in the surface film because they are utilizing the increased concentrations of organic matter or do they accumulate by the same mechanisms that concentrate DOC at the sea surface? One study suggested the probable mechanism was a concentration process rather than growth in the surface film. Kjelleberg & Håkansson (1977) isolated bacteria from the surface film and subsurface coastal water and assayed over 1000 isolates for their ability to grow on media containing lipid, protein, and carbohydrate. The results showed a statistically significant difference between the isolates from the surface film and the subsurface water but the activity of the surface film bacteria was always lower than that of the subsurface isolates. Although these results are based on the minority of bacteria which are amenable to culture, Kjelleberg & Håkansson could find no evidence that the accumulation of bacteria in the surface film was a result of growth. Also, in a study of the chemistry of the surface film, Wallace & Duce (1978) could not measure any conversion of DOC to POC and they suggested that bacteria in the surface film have very slow growth rates. Various processes have been proposed as the mechanism which accumulates material in the surface film but transport by bubbles is generally considered to be the most important. Blanchard (1964) showed that bubbles rising through a water column adsorb dissolved and particulate matter which has hydrophobic surfaces; the bubbles burst on reaching the surface but ejected only a small fraction of these organics into the atmosphere, the rest being left in the surface film. This process has also been suggested as a mechanism of concentrating bacteria (Carlucci & Williams, 1965). In laboratory experiments, Bezdek & Carlucci (1972), Blanchard & Syzdek (1974), and Norkrans & Sørensson (1977) showed that marine bacteria could attach to bubbles and be carried to the surface layer with an efficiency of about 0.1%. Tsyban (1971) commented that such a mechanism of bacterial accumulation in the surface film is apparently more efficient than the loss to subsurface water by turbulent diffusion because bacteria are even found in the surface film in stormy weather. Surface charge and hydrophobicity of the cell surface have been implicated in the interaction of bacteria in culture and lipid surface films (Kjelleberg & Stenström, 1980) and, in a recent study, Dahlbäck, Hermansson, Kjelleberg & Norkrans (1981) found a positive correlation between the degree of enrichment in the surface film and the hydrophobicity of the bacterium; hydrophobic bacteria would adhere better to rising

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bubbles and to the surface with a resulting concentration of bacteria in the surface layer. These data suggest that bacteria accumulate in the surface film by the same mechanisms that concentrate organic matter rather than as a result of growth. But, if there are high concentrations of organic matter, why do the bacteria not exploit this and grow rapidly? Obviously further experimentation is required to measure the metabolic rate of surface film bacteria but Kjelleberg & Håkansson (1977) speculate that the high intensity of ultraviolet light at the ocean surface may inhibit bacterial growth. DEEP-WATER COLUMNS Bacterial activity in deep water is controlled by a number of factors. In addition to the low availability of utilizable organic matter, bacteria must contend with extremely high pressure and low temperatures; 90% of the ocean is at a temperature of 5°C, or less, and pressure in the abyssal trenches is over 1000×105 Pa (1100 atm) (Morita, 1976). The effect of low temperature on bacterial activity has been the subject of several recent reviews (Morita & Haight, 1964; Morita, 1975, 1976; Baross & Morita, 1978) and the result of elevated pressure on bacterial activity has been reviewed by Morita (1967, 1972), Marquis (1976), Marquis & Matsumura (1978), and Landau & Pope (1980). Marine bacteria can be arbitrarily classified on the basis of their tolerance to pressure; some are relatively sensitive to pressure and are incapable of growth at pressures greater than 200×105 Pa (200 atm), some are relatively tolerant growing at pressures up to 500×105 Pa (500 atm) and others are highly barotolerant growing at pressures up to 1000×105 Pa (1000 atm) (Oppenheimer & ZoBell, 1952; Kriss, Mitskevich & Cherni, 1969). Barotolerance appears to be related to growth substrate (Kriss & Mitskevich, 1967); Marquis, Brown & Fenn (1971) found the barotolerance of Streptococcus faecalis in culture, increased from 200 to 550 atm when glucose replaced pyruvate as growth substrate. Absence of organic matter may also affect barotolerance and Novitsky & Morita (1978b) reported that starvation increased the barotolerance of a marine vibrio in culture. Concentrations of various salts alter barotolerance of bacteria, NaCl resulting in the greatest enhancement (Palmer & Albright, 1970; Albright & Henigman, 1971). Increased pressure reduces the pH range over which bacterial cultures can grow (Matsumura, Keller & Marquis, 1974) but this may not be a particularly relevant observation in view of the relatively constant pH of sea water. The biochemical process which has been suggested as the most sensitive to pressure is protein synthesis (Pope & Berger, 1973; Albright & Hardon, 1974; Swartz, Schwarz & Landau, 1974; Albright, 1975) and, in particular, the structure of the 30S submit of the ribosome (Pope, Smith, Swartz & Landau, 1975; Smith, Pope & Landau, 1975). Other cellular processes such as catabolism and ATP production (Matsumura & Marquis, 1977), membrane transport systems (Schen & Berger, 1974; Baross, Hanus & Morita, 1974), and the maintenance of protein structure and enzyme activity (Mohankumar & Berger, 1974) at high pressure all contribute to the barosensitivity of some bacteria (Marquis & Matsumura, 1978). Similarly, psychrophilic marine bacteria appear to be sensitive to increased temperature because of the thermal instability of their enzyme systems (Christian & Wiebe, 1974; Morita & Mathemeier, 1977). The relevance of much of this work to deep-sea bacteria could be questioned because it

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is based on experiments with bacteria in culture, many of which were not even isolated from the deep sea and were first pressurized in the laboratory (Morita, 1970). The problems of sampling bacteria from the deep sea and maintaining ambient temperature and pressure during transfer to the laboratory are considerable and increase in complexity if attempts are made to isolate and culture bacteria under in situ conditions. It is only recently that improved sampler design has enabled a start to be made on in situ measurements and on laboratory experiments with bacterial isolates which have not been decompressed or temperature shocked during sampling. The first design of a pressurized sampling and culturing vessel was described by Jannasch, Wirsen & Winget (1973) and subsequently modified for use at greater depth (Jannasch, Wirsen & Taylor, 1976). A later design allowed the collection of a 3-l water sample which was concentrated to about 13ml and re-inoculated into pressurized water containing 14C-labelled amino acids to obtain an estimate of bacterial activity (Jannasch & Wirsen, 1977a). Turnover of organic matter in deep water by bacteria The constancy of dissolved and particulate organic matter in the deep ocean led to the assumption that microbial activity was low (Menzel & Ryther, 1968); this was not, however, based on any experimental evidence, which was not sought until the famous Alvin accident. The research submarine, ALVIN, was accidentally sunk in 1540 m of water off Massachusetts and was not recovered for 10 months; on recovery, the crew’s lunch was found to be remarkably well preserved with no evidence of microbial degradation, even although it was saturated with sea water and had presumably been in contact with deep-sea microbes (Jannasch, Eimhjellan, Wirsen & Farmanfarmaian, 1971). The food quickly spoiled when kept at atmospheric pressure in a refrigerator, suggesting that the physical environment of the deep sea had inhibited microbial action, rather than that spoilage had been prevented by the presence of bacterial inhibitors. This chance observation of apparent lack of microbial activity was the stimulus required for more research work on the turnover of organic matter in the deep ocean (see reviews of Jannasch & Wirsen, 1977b; Jannasch, 1979b; Morita, 1979). Jannasch et al. (1971) incubated water samples taken from the surface and 200 m with four 14C-labelled substrates at a depth of 5300 m for 8 wk and compared the total uptake over that period with duplicate samples kept at 3 °C at atmospheric pressure for 6 wk. The uptake rate of the in situ samples, relative to the refrigerated controls, was 88 times slower for carbohydrate, 62 times slower for glutamate and 11 times slower for casamino acids. In a subsequent study, Jannasch & Wirsen (1973) repeated this experiment using indigenous bacteria from deep water and again found very low uptake rates which were between one and three orders of magnitude less than refrigerated controls at atmospheric pressure. Particulate matter, such as paper, wood, and Ulva thalli, showed no measurable degradation after a one year incubation in bottles at 1830 m. As a check that these experimental procedures were not producing artefacts, Wirsen & Jannasch (1974) used exactly the same procedures with surface sea-water samples; there did seem to be a methodological problem because in situ activity was only 50% of the laboratory control activity but the temperature difference of 3 °C might have accounted for this. Despite

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these methodological problems, the in situ reduction in activity was never as great as that recorded for the deep-sea samples and it is reasonable to accept the very low metabolic rates in the deep sea. Seki, Wada, Koike & Hattori (1974) took a different approach to studying the activity of deep-sea bacteria, involving the incubation of an in situ sample, enriched with the marine medium 2216E of ZoBell (1946), at 5200 m for 5 days; the numbers of bacteria which developed in the enriched cultures were then assessed by microscopy. The results suggested that deep-sea bacteria had a generation time of several days with an activity one to two orders of magnitude less than surface bacteria. Seki et al. (1974) also suggested that a small proportion of these barotolerant bacteria had the potential for very rapid growth if a suitable organic substrate were available and this would result in the rapid degradation of organic particles. Jannasch et al. (1976) did the first in situ timecourse experiment of uptake of labelled organics; they found an initial phase of uptake which ceased even although over 80% of the substrate remained unused. The authors suggested that hydrostatic pressure and low temperature were responsible, in some unexplained way, for this retardation of microbial activity. When the pressure in the samples was reduced to atmospheric and the temperature raised to room temperature, after a brief lag there was a sharp increase in incorporation and respiration. Jannasch et al. (1976) also commented that, if the definition of a barophile is a bacterium which grows better at high pressure than at atmospheric pressure, then the marine bacteria which they sampled did not appear to be barophilic, but merely barotolerant because there was never a stimulation of bacterial activity at elevated pressure. A different estimation of bacterial activity in the deep sea was made by Williams & Carlucci (1976) using oxygen consumption data obtained with a deep-sea isolate, capable of growth on natural low nutrient concentrations; they estimated the oxygen consumption in the deep sea to be 1.4 ml O2·m−3·yr−1 (equivalent to 0.61 mg C·m−3·yr−1). This compares with carbon utilization rates of 2–13 mg C·m−3·yr−1 obtained by Jannasch et al. (1971). Williams & Carlucci (1976) used this estimate of bacterial respiration to construct a carbon budget for north central Pacific at 2000 m. Assuming a steady state, they calculated the DOC residence time to be 738 years which is much less than the value of 2600 to 3500 years estimated from measurements of 14C natural abundance in the same region (Williams, Oeschger & Kinney, 1969). Williams & Carlucci (1976) also suggested that DOC must be the main source of energy for bacteria as POC utilized at the rate suggested from the oxygen consumption data, would be depleted within five years and result in a measurable gradient of POC from 500 to 2000m; such a gradient has not been found, although it is questionable whether existing sampling techniques are sufficiently sensitive to exclude the possibility of such a gradient. Using an approach based on the activity of enzymes of the respiratory electron transport chain, ETS activity (Curl & Sandberg, 1961; Packard, Healy & Richards, 1971), Christensen, Owens, Devol & Packard (1980) estimated the oxygen consumption rate at 5000 m in the eastern tropical North Pacific to be 5 ml O2·m−3·yr−1, i.e. greater than the estimate of Williams & Carlucci (1976); this estimate is probably too high as it could result in large areas of the ocean becoming anoxic. Therefore, bacterial activity in the deep ocean must be quite low with a slow rate of turnover of organic matter balancing the low supply of organic matter from surface water.

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Isolation of deep-sea bacteria The study of processes involved in the turnover of organic matter by bacteria in deep water has been hindered by the absence of cultures of deep-sea bacteria with which to experiment; the technology is now available for growing and experimenting with cultures under pressure and such research is likely to be very much less expensive than in situ studies and should greatly aid our understanding of the way in which bacteria respond to low organic concentrations at low temperatures and high pressure. The few deep-sea bacteria which have been isolated are barotolerant, rather than barophilic; that is, they do not grow better at the high pressures found in the deep-sea than at atmospheric pressure (Colwell & Kettling, 1974; Wirsen & Jannasch, 1975; Jannasch et al., 1976; Williams & Carlucci, 1976; Jannasch & Wirsen, 1977a). Yayanos, Dietz & Van Boxtel (1979), however, have reported the isolation from dead, deep-sea amphipods, of a truly barophilic bacterium, with a generation time of 4–13 h at a pressure equivalent to 5700 m depth but with a generation time of 3–4 days at atmospheric pressure. Unlike previous isolates of deep-sea bacteria which apparently lost their barotolerance on successive transfer in pure culture (Schwarz, Yayanos & Colwell, 1975), this isolate appears to be truly barophilic. The bacterium was isolated from a deepsea animal, and had access to high organic concentrations so it may not be very representative of free-living bacteria in the deep sea; nevertheless, the isolation and culture of a truly barophilic bacterium gives some hope that cultures may soon be established of low nutrient barophilic bacteria which can be used in experiments of organic turnover under deep-sea conditions of pressure and temperature. The rôle of deep-sea fauna in organic degradation The experiments carried out immediately after the recovery of the lunch-box from the ALVIN enclosed particulate organic matter in bottles which prevented any interaction with the fauna of the deep sea. Sieburth & Dietz (1974) suggested that the exclusion of omnivorous scavengers from the Alvin lunch-box might have disturbed the natural process of biodegradation and resulted in the preservation of the food. They duplicated the Alvin experiment by incubating foodstuffs at between 3000 and 5000 m in various types of enclosures, some of which totally excluded fauna and others allowed access by different size classes of animal. Food in perforated boxes was decayed within 10 wk in the deep sea but the food was in an excellent state of preservation in boxes which excluded animals. Wirsen & Jannasch (1976) planted perforated tubes filled with solid nutrient agar into deep-sea sediment and found that the agar above the sediment surface was attacked by small invertebrates, but the media below the surface was untouched by microbial or animal activity. These studies emphasize the importance of animal and bacterial interaction in the degradation of participate organic matter. It is a common observation that organic matter degradation in terrestrial soils is accelerated by the action of soil animals which prime plant material to produce a large surface area to be colonized by soil microbes (Gray, 1976; Lee, 1980). Clearly, similar mechanisms may be implicated in the

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sea; the degradation of organic matter by the deep-sea community is greater than that of a single isolated component, such as bacteria. Bacteria in the gut of deep-sea animals are also important in the overall transformation of organic matter of the sea. In a recent study of deep-sea amphipods, Jannasch, Cuhel, Wirsen & Taylor (1980) followed the in situ transformation of 14C-labelled fish meat; the amphipods converted a large fraction of the ingested food into storage products, such as lipid, but a significant fraction of 14C was found in nucleic acid. The authors argued that the amphipods were not growing sufficiently to produce a detectable increase in nucleic acid so the label incorporated into nucleic acid must be due to growth of bacteria in the gut; they suggested that, by looking at the distribution of label in nucleic acid and lipid, it should be possible to distinguish the metabolic activity of the amphipod from the gut flora.

BACTERIA IN DEEP-SEA SEDIMENTS The low metabolic activity of bacteria in the deep-sea water column is also reflected in the sediment. Early estimates of sediment activity were obtained by measuring oxygen consumption by cores of deep-sea sediment incubated on board ship (Pamatmat & Fenton, 1968; Pamatmat, 1971) but it was impossible to partition oxygen consumption between bacteria and animals in these studies, although chemical oxygen demand could be estimated by poisoning the sediment with formalin. Working in the eastern Pacific with cores from 2750m and 2900m, Pamatmat (1971) measured oxygen consumption rates of 1.5–4.5 ml O2·m−2·h−1, with chemical oxygen demand accounting for 66–106% of the total uptake. These rates are up to an order of magnitude greater than those in a later study by Smith & Teal (1973) who reported a mean uptake rate for ten incubations off New England, of 0.5 ml O2·m−2·h−1 using in situ bell-jar incubations; no chemical oxygen demand was detected in the presence of formalin. Smith & Teal (1973) speculated that the difference between their estimates and those of Pamatmat (1971) was a result of decompression of the sediment cores and increased supply of organic matter in the Peruvian upwelling region studied by Pamatmat. In a subsequent study, Smith (1974) attempted to assess the contribution of bacteria to total O2 uptake by in situ incubations with the antibiotics, streptomycin and penicillin; bacterial respiration was 8–8% and biological O2 demand was 54.6% of the total mean oxygen uptake of 2.4 ml O2·m−2·h−1. These measurements were made in the San Diego Trough which has a greater surface productivity than the region previously studied by Smith & Teal (1973) and probably accounts for the higher oxygen consumption rates. These estimates of the contribution of bacteria to total sediment respiration, however, appear too low; streptomycin and penicillin are bacteriostatic rather than bacteriocidal and only act effectively on actively growing bacteria. Considerable oxygen consumption may occur in the presence of these antibiotics if the bacteria are not rapidly dividing and metabolism reflects only maintenance requirements; in all probability, this is the situation in deep-sea sediments and antibiotics would only inhibit a small proportion of the total bacterial respiration. More recently, Smith et al. (1976) described the design of an improved respirometer which they deployed in the northwestern Atlantic; they compared respirometer

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measurements with various other methods including a grab respirometer which was retrieved after in situ incubation for analysis of the faunal constituents of the sediment sample (Smith, 1978; Smith, White, Laver & Haugsness, 1978). The numerically dominant animals in the sediment were small, being mostly nematodes and polychaetes, 75% of which passed through 420-µm sieve, clearly demonstrating the potential importance of meiofauna in the turnover of carbon in deep-sea sediments. Smith (1978) obtained predictive regression equations for benthic oxygen consumption; strong correlations were found between benthic community respiration and water depth but less significant correlations were found with water temperature, dissolved oxygen, benthic animal biomass, sediment organic matter, and surface primary production. Smith calculated that benthic respiration accounted for 15–29% of the surface primary production which actually reaches the bottom i.e. at 2200 and 3650 m benthic respiration utilized 1–2% of the surface primary production. Rowe, Pollini & Horner (1974) found an exponential decline in faunal density with water depth, which is in agreement with the findings of Smith (1978). Christensen & Packard (1977) estimated sediment oxygen consumption rate of 0.2 ml O2·m−2·h−1 at a depth of 1800 m off North West Africa, which is comparable to the estimates of Smith & Teal (1973). It is only recently that measurements have been made of oxygen concentrations in deep-sea sediments; previously the presence of oxygen had been inferred from the distribution of nitrate. Murray & Grundmanis (1980) measured oxygen concentration in pore waters of deep-sea sediments and confirmed that oxygen was present throughout the top 50 cm of an equatorial red clay and carbonate ooze in the Pacific. These authors found a decrease in oxygen concentration within the top 5 to 10 cm of most sediments but, below that depth, there was a constant, measurable concentration which was never less than 50 µmole·kg−1. Murray & Grundmanis (1980) measured a diffusive flux into the sediment of 0.08 ml O2·m−2·h−1 which is similar to the estimates of total, i.e. chemical and biological, oxygen demand made by Smith & Teal (1973) and Smith (1978). There have been relatively few studies which have measured bacterial activity, per se, in deep-sea sediments. Schwarz & Colwell (1975) obtained sediment samples from 7750 and 8130 m in the Puerto Rican Trench but their sampling procedure could not avoid decompressing the sediment samples during recovery; the samples were subsequently repressurized and incubated at in situ temperature and pressure. At high pressure, there was a diversion of carbon away from assimilation which resulted in a 22% increase in respiration of assimilated carbon; total activity was low and the uptake of 14C-amino acids was only 2% of the rate measured at atmospheric pressure. Other bacteriological studies have attempted to investigate the ability of bacterial isolates to grow on media of different organic composition (Morita & ZoBell, 1955; ZoBell & Morita, 1957, 1959; Quigley & Colwell, 1968; Hogan & Colwell, 1969; Colwell & Kettling, 1974; Norkrans & Stehn, 1978) but in none of these studies was an attempt made to prevent decompression of the sediment during sampling; it is probable that bacteria isolated from deep-sea sediment, exposed to atmospheric pressure and which can be cultured on high nutrient media, will be very unrepresentative of the bacterial flora of deep-sea sediments. HYDROTHERMAL VENTS IN THE DEEP SEA

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An interesting example of a special environment in the deep sea which appears to have been successfully exploited by bacteria, occurs at hydrothermal vents in regions of seafloor spreading and recent vulcanism (Francheteau et al., 1979; Speiss et al., 1980; Edmond, 1981). Lonsdale (1977) first described extensive communities of large animals near to these hydrothermal vents; many of the animals had never before been described and they grew to remarkable sizes. The obvious question was, what were these very large animals eating and why was their distribution restricted to the hydrothermal vents? Jannasch (1979a) and Jannasch & Wirsen (1980) suggested that chemoautotrophic bacteria, utilizing the reduced sulphur emitted from the vents, might be a novel food source for these animals. Very large numbers of bacteria were initially reported in water as it flowed from the vents with cell densities approaching that attainable in culture, i.e. 108–109 bacteria·ml−1 (Corliss et al., 1979). Karl, Wirsen & Jannasch (1980) subsequently reported, however, much lower cell densities of 5×105–106 bacteria·ml−1. ATP measurements confirmed these lower biomass estimates which were reduced from the 1 g C·m−3 originally estimated by Corliss et al. (1979) to 100–250 mg C·m−3; this still indicates considerable bacterial production. Although Karl, Wirsen & Jannasch (1980) attempted to measure chemoautotrophic CO2 fixation, their results were equivocal but the rates were higher than had previously been found in the Black Sea, where chemoautotrophy is implicated (Tuttle & Jannasch, 1973). Additional evidence for the chemoautotrophic source of carbon in these regions came from an analysis of 13C/12C ratios in a hydrothermal mussel. Rau & Hedges (1979) reported that the tissues of the mussel were depleted in 13C, when compared with other marine animals, and this would be consistent with the animal feeding on chemoautotrophic bacteria, and not carbon derived from photosynthetic primary production. Enright, Newman, Hesser & McGowan (1981) have, however, recently challenged the hypothesis that chemoautotrophs are responsible for maintaining the large faunal biomass at hydrothermal vents. They pointed out that the temperature of the dominant water flow issuing from the vents is over 300 °C and much too high for any bacteria to survive; also many of the larger animals are found many metres away from the vents and it is difficult to believe that they would encounter significant concentrations of bacteria which might emanate from the vents. Enright et al. (1981) considered other possible sources of food for the fauna and suggested that deep-water organic matter, initially derived from photosynthesis, could be advected into the region of hydrothermal activity as a result of water currents produced by the rising plume of warm, buoyant, vent water. They suggested that such a process could supply up to 30 kg of paniculate organic matter per day. This should, however, result in the accumulation of organic matter near the vents because any particles carried upwards in the rising, warm water should be quickly redeposited within the region of lateral water movement and returned to the base of the rising plume. Such a mechanism would result in animals living near the vent receiving an increased concentration of organic matter, relative to other deep-sea animals. It is, however, difficult to distinguish experimentally between the two hypothesis. Plumes of buoyant water at hydrothermal vents would also accumulate any chemoautotrophic bacteria growing in vent water and so chemical analysis could not distinguish between

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the sources of supply. Further work is obviously required to show that chemoautotrophic bacteria are abundant at hydrothermal vents and that their growth rate is sufficient to maintain the observed animal biomass. Only then will it be possible to say unequivocally, that chemoautotrophic bacteria can be an important alternative source of primary production in certain specialized deep-sea environments. THIOPLOCA Incredibly dense microbial biomass has been reported in the sediments of the upwelling region off Peru and the main component is the filamentous bacterium, Thioploca (Gallardo, 1977); he reported biomass densities of 106 g wet wt·0.1 m−2 sediment which gave the surface sediment a soft spongy texture. Thioploca appears to be a useful, if not the major, food source for higher organisms in this region. Thioploca is classified as a member of the Beggiatoacea (Leadbetter, 1974) but recently, Morita, Iturriago & Gallardo (1981) questioned the assumption that reduced sulphur is the energy source for Thioploca growth and suggested that methane was both the source of energy and of carbon for growth; the postulated source of methane was production by methanogens in the sediment and seepage out of coal deposits in the region. Thioploca did not assimilate acetate, glucose or amino acids and Morita et al. (1981), therefore, considered Thioploca to be a methylotroph, whose growth on methane resulted in a significant supply of organic matter to the sediments of the Peru-Chile upwelling system.

CONCLUSION Marine bacteria have traditionally been considered as mineralizers and other consequences of bacterial growth have usually been ignored. A gradual reassessment, however, is taking place of the rôle of bacteria in the sea. The assumed close coupling of organic matter transfer from phytoplankton to zooplankton is being questioned and may not be valid; a considerable proportion of primary production may pass into the DOC and POC pool and it is reasonable to expect bacteria to utilize this organic matter. But, is the rôle of bacteria restricted to mineralization, a convenient way of removing ‘waste’ organic compounds from an ecosystem or do bacteria make a significant contribution to the recovery of this organic matter and its re-introduction into the food web as a result of bacterial production and subsequent grazing by other trophic levels? In a recent, thoughtful review. Williams (1981) considered the place of bacteria in the planktonic food web and concluded that “microheterotrophs may pass on twice as much organic matter to the next trophic level as herbivorous zooplankton”. As more information is obtained on the activity of marine bacteria, it is probable that a fundamental reassessment of the rôle of bacteria in the sea will be necessary.

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PARTICULATE MATTER IN THE OCEANS— SAMPLING METHODS, CONCENTRATION, SIZE DISTRIBUTION AND PARTICLE DYNAMICS W.R.SIMPSON Institute of Oceanographic Sciences Brook Road, Wormley, Godalming, Surrey, U.K.

Oceanogr. Mar. Biol. Ann. Rev., 1982. 20, 119–172 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION The suspended particulate matter controls the chemistry of the oceans; the more reactive an element or chemical species the more rapidly it is removed from solution to the particle phase and the faster it sediments. There are many complex interrelated processes to be considered in the elucidation of particulate and element cycling. For example, the three dimensional relationship of latitude, longitude, and depth compared with particle concentration and composition needs to be known in the context of physical variables such as temperature, salinity, and density; adsorption, absorption, desorption, and dissolution require estimation for different particle types and chemical species; the biological involvement in terms of uptake, excretion, and settling of tests plays a major rôle as does the physical aggregation, fragmentation, and disaggregation of such particles. Suspended particulate matter in the oceans has four main sources: (i) fluvial input of terrigenic material; (ii) aeolian input from wind erosion of continental masses, volcanism and man-made sources; (iii) resuspension of sedimentary material by current erosion, earthquakes and slumping; and (iv) authigenic production by biota, submarine volcanism and precipitation of inorganic minerals. Of minor importance is the extra-terrestrial supply in the form of cosmic spherules. In the past there has been one over-riding difficulty in the study of such processes, that of the low concentration of particulate matter in sea water, typically ≈10 µg·1−1. Over the last two decades, improvements in sampling methods, new instrumentation, and better

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micro-analytical techniques have made possible significant advances in our understanding of oceanic particle dynamics, exemplified by the Geochemical Oceans Sections Study (GEOSECS, 1970, 1974, 1976, 1980; Lal, 1977). This does not mean that we now have all the answers but it has raised a second generation of problems and questions. The purpose of this article is to review existing sampling methods, to discuss data on particle concentration, size distribution, fluxes, composition and cycling processes, and to highlight areas for improvement of methodology and the needs of future research. Excellent reviews of the subject have been written by Lal (1977) and Sackett (1978); the theory of geochemical processes presented by Lerman (1979) is invaluable and there is a comprehensive collection of papers edited by Gibbs (1974) on suspended solids in sea water. The review is arranged so that the procedures which lead to general information on the horizontal and vertical distribution of mass concentration and particle size spectra are discussed (pp. 120–136). Sampling procedures that are used at present to supply information on fluxes and turnover either directly (sediment traps) or indirectly (in situ filtration systems) are then examined (pp. 136–145). The theory of particle dynamics that has found practical application (or may do so in the future) is dealt with on pp. 145–155. The theory is supported by field observations and some ways of over-coming experimental difficulties are discussed. One approach that provides valuable information on the net rôle of particles in scavenging and removal processes is the measurement of radionuclides, in both particulate and solution phases (see pp. 155–166). Considered outside the sphere of this review is any detailed discussion on the bulk stable element composition of particles.

THE MEASUREMENT OF PARTICLE CONCENTRATION Gravimetric analysis, after filtration or centrifugation, has been the classical method for the determination of particle concentration and is still used today although this method is likely to be superseded by optical methods of either light scattering (nephelometry) or light absorption (transmissometry) that offer the considerable advantage of continuous monitoring. GRAVIMETRIC ANALYSIS Water bottle samples The total concentration of particulate matter in the oceans is so low that large volumes of water are required to obtain quantifiable samples. Generally, 30-l Niskin samples are employed followed by pressure or vacuum filtration of the sample. There are two main sources of error concerned with water bottle sampling. First, where bottom spigots are side mounted, large, rapidly settling particles may settle below the spigot outlet and, unless the necessary precautions are taken, remain in the bottle dregs. This may produce an underestimate of concentration particularly for surface and bottom waters (Bishop &

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Edmond, 1976; Gardner, 1977; Calvert & McCartney, 1979). Secondly, the volume sampled may not include relatively rare large particles estimated to be at a concentration of two 100-µm particles per litre and one 1-mm particle per 50 to 500 litres (Sheldon, Prakash & Sutcliffe, 1972), yet these particles supply most of the mass flux to sediments. To illustrate this point, Figures 1 and 2 show scanning electronmicrographs of Rockall Trough material (1610 m) collected by water bottle and sediment trap. (Nott & Simpson, unpubl.).

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Fig. 1.—Scanning electron micrographs of suspended material from the Rockall Trough (Nott & Simpson, unpubl.): variation in quantity and type of material with depth; samples were collected in 7–1 N.I.O.

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bottles at (a) 35 m, (b) 510 m, (c) 1100 m, and (d) 1610 m.

Fig. 2.—Scanning electron micrograph of material collected in a sediment trap in the Rockall Trough after a 9-day deployment at 1610 m: note the variation in the quantity of material compared with Figure 1(d); scale as in Fig. 1.

Figure 1 also illustrates the change in particle type and concentration with depth. All samples were filtered onto Nuclepore membranes. The precision in weighing samples on such filters after the filtration of 81 of water was put at ±1.0 µg·1−1 by Betzer, Carder & Eggiman (1974). The error includes the accuracy of the weighing, retention of salt, and variation in desiccation efficiency. The effect of different filters is discussed on pages 142–144. Continuous centrifugation The main limitation of continuous centrifugation is the depth from which water can easily be pumped to the ship unless large water bottles can be used, e.g., 2001 (Jacobs & Ewing, 1969). Bassin, Harris & Bouma (1972) suggested that the centrifugal force of 9000g, as used by Jacobs & Ewing (1969), was inadequate to spin down all clay and organic particles; forces of up to 100000g may be required, which are impractical for ship work. The most comprehensive data for the world oceans were based on this method prior to 1970 (Lisitzin & Glazunov, 1960; Jacobs & Ewing, 1969; see Table I) but Jacobs & Ewing (1969) found significant variations compared with results given by Lisitzin

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(1960). They also found contamination a problem and even their data may be high by a factor or two or three (Manheim, Meade & Bond, 1970). OPTICAL METHODS The scattering of light is dependent on the size distribution, concentration, refractive index, and shape of the particles with most of the scattering occurring at low angles in the forward direction. The Mie theory (Mie, 1908) is used in the calibration of nephelometers but, because of the variables mentioned above, the results are semi-quantitative (Sackett, 1978). Discussions of some of the variables are given by Gordon (1974; Mie-theory models in relation to size-refractive index distributions), Smith, Austin & Petzold (1974; volume-scattering functions), and problems encountered in beam transmissometry by Tyler, Austin & Petzold (1974). Nephelometers Nephelometers at present in use are based on two original designs, one from the LamontDoherty Geographical Observatory (Thorndike & Ewing, 1967) and the other from the University of Washington (Ahlquist & Charlson, 1968). The former consists of a lightsource-baffle and calibrating attenuator-camera arrangement. A constantly moving 35 mm film records scatter of angles between 8° and 30° as bands of light on either side of the centre of the film. The profile is read from time marks on the film and this can be related to ‘wire out’. Measurements are recorded as log scattering (log S) where log S= (log of the exposure of scattered light) minus (log of the exposure of the attenuated light). The University of Washington instrument was adapted by Sternberg et al. (1974) to incorporate a xenon flash tube and photomultiplier mutually at right angles; the photomultiplier measured the scattered light intensity of angles between 10° and 170°, the signal being recorded on an internal chart strip. Calibration to convert scattering to concentration (µg·1−1) was carried out using gravimetric data of the inorganic component after filtration (correlation coefficient, r=0.73). This is the normal practice for calibration; other correlations reported are 0.91 (Biscaye & Eittreim, 1974), 0.62 (Bassin, 1975), and 0.85 (Feely, 1975). Further modifications to this model were introduced by Bassin (1975) of Texas A & M University and both the LDGO and TAM instruments were used by Feely (1974) in the Gulf of Mexico. Diagrammatic representations of the instruments are given in Figure 3a and b. Transmissometers Sackett (1978) quotes the detection limits for transmissometers of 0.5 mg·1−1 compared with 0.5 µg·1−1 for nephelometers. There is, however, now a commercially available instrument, based on the Oregon State University design (Bartz, Zaneveld & Pak, 1978) which offers resolution of 3 µg·1−1 and an

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Fig. 3.—Optical instruments for the continuous profiling of particulate concentration: a, Lamont-Doherty nephelometer; b, TAMU nephelometer; c, Sea-Tech transmissometer; a and b after Feely (1974).

operational depth of 7500 m. The 1-m light-beam, folded by a porro prism, operates at a wavelength of 660 nm to avoid attenuation due to yellow (organic) matter in the water and hence the attenuation of the suspended solids and water only are measured. The light source is modulated and synchronous with the detector. The small ratio of beam diameter (16 mm) to path length (1 m) serves to reduce the spurious flux derived from small angle forward scattering (Latimer, 1972). A diagrammatic representation of the instrument is

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shown in Figure 3c. Absolute calibration, again, is done gravimetrically by the procedure described by Bartz et al. (1978); performance with turbidity standards is given by Zaneveld, Spinrad & Bartz (1980). DISTRIBUTION OF PARTICLE CONCENTRATION The historical developments from the pioneering light scattering work of Kalle (1939) and Jerlov (1953) onwards are dealt with in detail by Sackett (1978). It is not intended to repeat his analysis here but to discuss results which describe the general trends observed in the oceans. Jerlov (1953) must be accredited with the first description of the general vertical particle distribution. Thorndike & Ewing (1967), however, succinctly described typical particle profiles in their early work with the photographic nephelometer. “The films show wide variations in light scattering with location. Definite trends and patterns are, however, evident: (1) surface water normally has high and variable light scattering; (2) in the open ocean, mid-water tends to have low scattering; (3) the variation of light scattering with depth is often smooth and gradual but occasionally it shows abrupt changes; (4) in some regions light scattering increases near bottom. The increase may be large or small, abrupt or gradual.” The first point (1) is due to biological productivity and terrestrial input and the second (2) describes the “clear-water” found at mid-depths. The existence of a nepheloid layer (4) was attributed to either transport of particles in the water mass from their source or to the resuspension of sediment (Ewing & Thorndike, 1965). Both are now known to be true, e.g., different water masses in a water column may be identified from their particle concentration and chemical composition. Maurice Ewing and his co-workers then went on to show the presence of the nepheloid layer in the northeastern Atlantic (Jones, Ewing, Ewing & Eittreim, 1970), northwestern Atlantic (Eittreim, Ewing & Thorndike, 1969), Arctic (Hunkins, Thorndike & Mathieu, 1969), Indian-Pacific Antarctic (Eittreim et al., 1972), and Argentine Basin (Ewing, Ewing & Le Pichon, 1967; Ewing, Eittreim, Ewing & Le Pichon, 1970). The time and spatial variability of the nepheloid layer, dependent on the water movement, has since been studied (e.g., Biscaye & Eittreim, 1974; Feely, 1974; Baker, 1976; Richardson, 1980). Transmissometer data is available for the coastal southern Pacific (Pak, Menzies & Zaneveld, 1979) and the Congo River and Angola Basin (Zaneveld, Spinrad & Menzies, 1979). Table I gives some indication of the variation of particle concentration in the oceans (Jacobs & Ewing, 1969). Although the overall values are somewhat greater than more recent data the same trends are observed, e.g., Atlantic concentrations are two to three times greater than Pacific concentrations where resuspension is low (Lal, 1977). Early studies on the Atlantic are

TABLE I Summary of weight-concentration data for suspended paniculate matter for whole water column in the major oceans (after Jacobs & Ewing, 1969, in Sackett, 1978)

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Range (µg·1−1)

Number of samples

Mean (µg·1−1)

North Atlantic

88

0.5–247

49

South Atlantic

52

1.5–197

51

North Pacific

145

0.5–152

37

South Pacific

78

4.5–86

30

Indian

37

9–177

72

Caribbean

25

0.5–139

40

Gulf of Mexico

11

12.5–193

66

summarized in Table II (after Manheim et al., 1970) and illustrate the trend to lower concentrations over the years as the methodology has improved. The gravimetric data collected during the GEOSECS work (Fig. 4) confirm the influence of the proximity of land, productivity, bottom currents, and water masses on particle distribution and concentration (Brewer et al., 1976). This work indicates a concentration range of 5–300 µg·kg−1 in the Atlantic with high concentrations associated with the Denmark Strait overflow, Antarctic Bottom Water and a plume seen between 35° N and 40° N due to sinking Labrador Sea Water. Lowest concentrations, 12 µg·kg−1, appeared in the midwater regions of the tropical gyres. Similar maps, on a smaller scale were constructed by Pak (1974) for the East Equatorial Pacific. Latitudinal variations in concentration for Atlantic surface waters were described by Krishnaswami & Lal (1977) and Lal (1977) with highest concentrations at extreme latitudes and lowest values between 40° N and 40° S (Fig. 5). At latitudes >40° S the material consisted mostly of biogenic opal yet calcium carbonate levels were nearly uniform with latitude at 10 µg·kg−1. The detrital composition varied from 0.3 to 9%, the highest concentrations appearing in glacial and Trade Wind regions. The average organic content was about 25 %.

PARTICLE SIZE DISTRIBUTIONS A knowledge of the particle size distribution and the type of particles that make up each size interval is essential, for example, in the calculation of fluxes, understanding processes, the description of plankton communities or as a variable in the scattering of light. Of the many methods available (see Swift, Schubel & Sheldon, 1972), only impedance, light blockage, and scanning electron microscopy and the associated difficulties will be discussed here.

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Fig. 4.—Suspended paniculate material in the western Atlantic: the most striking features are the high concentrations in near surface and near bottom waters (after Brewer et al., 1976); figures on contours are suspended particulate matter in µg·kg−1.

TABLE II Total suspended sediment concentrations in open Atlantic waters: membrane filters refer to filters with pore size of 1 µm or less; after Manheim, Meade & Bond (1970)

Area

Depth

Dominant conc, range (mg·1−1)

Remarks

Source

Mid-Atlantic Ridge Surface about 52° N

0.32 (SiO2)

Single sample; paper filter

Murray & Irvine, 1891

North Atlantic

Variable

0.05–1.0

Membrane filter

Armstrong, 1958

Cape Farewell to Flemish Cap

Variable

0.02–15

Paper filter

Krey et al., 1959

Northern North Atlantic

Surface

0.5 (av.)

1.8 µm filter

Krey, 1964

Deeper water

0.1 (av.)

North and South Atlantic

Variable

0.1–1.0

Membrane filter; higher conc, near bottom

Klenova et al., 1962

North Atlantic

Variable

0.3–3.0

Membrane filter

Vikhrenko & Nikolayeva, 1962

East of Blake Plateau

4030m

2.5

Single sample; Groot & Ewing, continuous centrifuge 1963

Eastern North and

Surface

0.7 (av.)

Membrane filter

Gordeyev, 1963

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South Atlantic 0.1–0.3

Continuous centrifuge Gordeyev, 1963

Western North Atlantic

Surface

0.2–1.0

Membrane filter

Vikhrenko, 1964

Central Gulf Stream

Surface

0.13

Membrane filter

Krey, 1961

Deeper water

0.06 Membrane filter

Hagmeier, 1964

Tropical Atlantic

Surface

0.1–0.2

Deeper water

0.02–0.08

Tropical Atlantic

Variable

0.02–1.0

Membrane filter

Klenova & Vikhrenko, 1965

North Atlantic

Just below surface

0.04–0.14

Membrane filter

Folger & Heezen, 1968

Subtropical western North Atlantic

Variable

0.001–0.25 (mean 0.05)

Continuous centrifuge Jacobs & Ewing, 1969

Western North Atlantic

Surface

0.1

Membrane filter

Manheim et al., 1970

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Fig. 5.—Variation of surface participate material with latitude in the Atlantic: a, total concentration; b, particulate organic carbon; c, ash content; after Lal, (1977).

IMPEDANCE-TYPE COUNTERS The Coulter counter operates on an impedance principle whereby a particle suspended in sea water, the electrolyte, passes through an aperture thus causing a drop in current across the aperture equivalent to its volume. The equivalent diameter is computed from the volume. The relationship is linear for particles having diameters between 2 and 40% (on new instruments 1.5–60%) of the diameter of the aperture; the full range of size determined by various sensors is 0.6–800 µm. Dilution of the sample with filtered sea water may be necessary for surface and coastal waters; typical volumes taken for analysis range from 0–5 to 10 ml. Studies on oceanic particulates using this method have been made by Parsons & Seki (1969), Bader (1970), Brun-Cottan (1971, 1976), Sheldon et al. (1972), Betzer et al. (1974), McCave (1975), Kitchen, Menzies, Pak & Zaneveld (1975), Eisma & Gieskes (1977), Lerman, Carder & Betzer (1977), Eisma & Kalf (1979), and Richardson (1980). In situ impedance counters have been described for surface work with nets (Maddux & Kanwisher, 1965; Boyd & Johnson, 1969) and for the examination of particle layering at 10 cm resolution by allowing the instrument package to free-fall to 100 m (Brown, 1977). COUNTERS USING LIGHT The HIAC cell uses a light-blockage principle whereby a particle passing through a collimated light beam in the sensor reduces the amount of light reaching a photodiode. The photodetector produces a voltage pulse of height directly proportional to the projected area of the particle. The dynamic range of the cell is 1:60 and has the advantage of continuous flow-through sampling. Pugh (1978) employed such a cell to gain information on plankton distribution in surface waters. A deep-water in situ counter (to 6000 m) that transmits data for the size range 10–600 µm in real-time with a sampling rate of 60 l·h−1 is under trial (Gwilliam, Lawford & Simpson, in prep.). It is hoped that in situ counting will reduce errors from fragmentation and aggregation incurred by sample handling and by the resuspension of filter and trap samples prior to analysis. McCave (pers. comm.) has recently found significant differences between Coulter and HIAC data for the range 1–20 µm. The charging of particles and/or the scattering of light are probably the sources of error, although noise cannot be discounted. For particles >500 µm, an in situ photographic instrument is under trial at the Woods Hole Oceanographic Institution; a ‘block’ of water of a known illuminated volume is photographed and particles measured and counted on the resulting prints (Honjo, pers. comm.). By this method, it is possible to gain some insight into the form and behaviour of the elusive ‘marine snow’. SCANNING ELECTRON MICROSCOPY (S.E.M.)

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Scanning electron microscopy (S.E.M.) was used by Harris (1977) for the examination of particles in the range 0.02–8 µm (although 0.45-µm pore size filters were used). Any microscopic method is laborious; Harris (1977) measured with dividers between 15000 and 45000 particles on enlarged prints (20×20 in). In this instance, however, the count was biased by a pretreatment that may have reduced the organic content by ≤20%. Baker, Feely & Takahashi (1979) also used S.E.M. for size analysis but at the magnification employed (×1000) there was an uncertainty in counts for particles <1.5 µm. The strength of S.E.M. is in the semi-quantification of the biological, mineral, and chemical composition of samples with respect to size and density. S.E.M. alone has been used to subdivide particle size into diatoms, coccoliths, aggregates, and unidentifiable fragments (Baker, Feely & Takahashi, 1979); to describe faecal pellets and faecal material (Bishop et al., 1977; Turner, 1977; Bishop, Ketten & Edmond, 1978; Honjo, 1978, 1980; Honjo & Roman, 1978; Emery & Honjo, 1979); to illustrate the dissolution of calcium carbonate (Honjo & Erez, 1978), foraminiferans (Corliss & Honjo, in press; Thunell & Honjo, 1981), and silica (Schrader, 1971); and to examine the distribution of surface particulates (Krishnaswami, Lal & Somayajulu, 1976; Emery & Honjo, 1979). When coupled with an E.D.X-R. system, S.E.M. work becomes central to the study of particle dynamics, adding a new perspective not obtainable by any other method to the biogeochemistry of ocean particles. Bassin (1974) identified suspended marine clays by this method and Jedwab (1979, 1980) applied the technique to the relatively rare mineral particles sampled during the GEOSECS experiments in the Pacific and Atlantic. Particles, both anthropogenic and natural in origin, were found to contain copper, zinc, lead (Jedwab, 1979), and many other elements. Dehairs, Chesselet & Jedwab (1980) showed barite to be a common component of suspended particulate matter; it was mostly biogenic in origin and of variable Sr/Ba ratios with evidence of dissolution at depth. The size of the particles ranged from ≈ 0–2 to 3 µm. Examples of analysis of faecal pellets, clays, glass shards, and sand grains have been presented by Honjo (1978) and Richardson (1980). Lambert et al., (1981) counted particles in the range 0.2 to 10 µm with a magnification of 5000×. For particles of a specific type, e.g., aluminosilicates, repeated 100 particle counts for given filters confirmed the reliability of the method. These same authors also counted opal debris, quartz grains, various types of aggregates, organics, and goethites (FeOOH). The distributions are discussed on pages 132–133. Simpson, Lane & Lunn (m prep.) used S.E.M.– E.D.X-R. analysis to subdivide particulates into seven main groups, biogenic calcium, biogenic silica, organic matter, faecal pellets, clays, glass shards and sand grains, and minerals on a frequency versus diameter basis. Examples of surface water and deep-water spectra are shown in Figure 6. Organic matter is seen to be prevalent in surface water but virtually absent from deep water. The frequency of occurrence of aluminosilicates increases with depth. Of the minerals, barite was common and iron was found in various forms such as coatings on biogenic particles, as precipitates, in clays, and as a component in aggregates. It is also possible to estimate the density of a particle or aggregate. Dry densities of the main forms encountered are given by Weast (1979) to be: opal 1.73–2.19, clays 2–3.3, calcite 2.72–2.94, aragonite 2.94, glass shards and sand grains (quartz) 2.75, iron oxides

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4.05–5.26, titanium oxides 4.2–5.5, and barite 4.5 (g·cm−3). The in situ densities of particulate matter, however were

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Fig. 6.—Illustration of the potential of S.E.M.-E.D.X-R. analysis in providing information on particle composition and distribution (Simpson, Lunn & Lane, in prep.): samples from the Cape Basin, 36°10′ S:9°1′ E; number size distribution with respect to particle type; note change in composition between depth 10 m (a) and 5000 m (b).

given as 1.6g·cm−3 for near bottom samples decreasing to values <1.0 g·cm−3 above the nepheloid layer (Gardner, Hollister, Spencer & Brewer, 1976). Organic matter is assumed to have a density close to that of sea water

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Fig. 7.—As Figure 6: showing the mass distribution of the particles in Figure 6 based on density difference; a, depth 10 m; b, depth 5000 m.

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(1–02) and the density of aggregates must be estimated from their composition. The mass distributions based on dry densities for particles in Figure 6 are given in Figure 7. Further discussions on the use of the S.E.M.-E.D.X-R. system will be entered into below but, as a semi-quantitative method, it shows great potential for the calculation of settling velocities and flux for particular particle phases as demonstrated by Lambert et al. (1981). PARTICLE DISTRIBUTION SPECTRA Data on particle size distributions in the water column most frequently approximate to the power-law relation

(1) where N is the cumulative number of particles (or particles per unit volume) greater than size d (which must be stated, usually 1 µm), A is the intercept, d is the equivalent sphere diameter (µm), and b is the gradient characteristic of the distribution, (Bader, 1970; McCave, 1975). Alternative distributions are discussed at length by Lerman (1979). The equivalent particle-volume distribution is

(2) and mass distribution

(3) where p is the density of the particles making up the distribution. The continuous function (Equation 1) may be written as

(4) Figure 8 gives examples of cumulative number distributions for the samples in Figure 6. A summary of data for the cumulative number distributions of suspended particulates is given in Table III. There is a trend in that the gradient, b, varies with size; typically b=2 to 4 for the range 0.04 to 4 µm, b=4 for 4 to 20 µm, and b=4 to 5 for particles or at least biogenic particles >20 µm. It is possible that the steepening of the gradient above 20 µm (e.g. McCave, 1975) is an artefact of the methodology, i.e., the small volumes taken for Coulter analysis do not include relatively rare larger particles. In most instances, the value of A decreases with depth until the nepheloid layer is reached. This suggests that

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630

large particles fragment or disaggregate with depth and that the smaller fragments dissolve or are removed from the system by some other means so as to produce the shallower gradient, b, at <4 µm. Although the power-law relation has found common usage to describe particle distributions, from the S.E.M. data of Lambert et al. (1981) there is evidence to show that small particles (<10 µm) of various types follow log-normal distributions. The point of inflection is at 0.8 µm just below the limit of functional Coulter counter analysis. The distributions are given by

(5) where D* is the diameter of a circle having the same projected area as any irregular particle measured, σ is the standard deviation of u, determined from the Laplace-Gauss function described in the original paper. Thus, it is apparent that there are few data on large particles including marine ‘snow’ yet faecal matter and large aggregates (≤1 cm) follow a power law distribution (Bishop et al., 1977, 1978, 1980).

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Fig. 8.—As Figure 6: showing cumulative number distribution of Figure 6 with respect to particle type; a, depth 10 m; b, depth 5000 m.

TABLE III Calculated values of the gradient, b, for various size ranges with depth in the oceans: some papers reported radius rather than diameter, in which case the radius values quoted have been doubled

Sea or Ocean

General distributions

Depth (m)

No. samples

Range of diameter (µm)

Mean gradient, b (range or ± SD)

Reference

Interlinking of physical

Mediterranean

300–800

E. Atlantic (Guinea Basin)

300–800

E. Pacific (Equatorial)

1000

200

632

1.4–20*

4.2(3.5–4.6)

Brun-Cottan, 1976

1.4–20*

4.1 (3.6–4.4)

”

3.23 (2.82– 3.43)

Carder, Beardsley & Pak, 1971

11 2.2–10.6

2 at 2000 N.W. Atlantic

Gulf of Mexico

N. Atlantic (Equatorial)

E. Pacific (Equatorial)

5–100

in McCave 1975 4 1–4*

2.42 (2.31– 2.64)

4–20

3.12 (2.88– 3.65)

200–4500

6 1–50

2.65 (2.39– 2.79)

600–3600

6 0.04–4†

‘2’

4–16

3.44

29–4522

15

4.08 ±0.32

30–2730

15

30–5111

14

30–900

8

3–5100

10

0–4940

12

4.06±0.53

0–4420

10

3.81±0.51

2–12

Sheldon in McCave, 1975

Harris, 1977

Lerman, Carder & Betzer, 1977

4.04+0.2 3.84±0.23 4.13±0.33

3−7.5*†

4.26±0.34

Baker et al., 1979

Selected particles Indian Ocean

3000–5200

4 0.2>50

4.5±0.3

Lal & Lerman, 1975

4 55–183

5.7 (4.6–7.7)

Bishop, 1977

Foraminifera fragments

64–146

5.3 (4.1–6.3)

Bishop et al., 1978

Faecal pellets

82–292

3.8 (2.8–4.6)

Foraminifera and diatoms S. Atlantic (Cape 20–400 Basin) Foraminifera

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Faecal matter E. Pacific (Panama Basin)

159

91–3652

4.3 (3.8–4.9)

Foraminifera

50–1000

8.0 (4.37– 11.04)

Faecal pellets

50–1000

3.0(1.80–4.60)

Faecal matter

50–10000

2.3(1.78–3.19)

15–1500

12 Bishop et al., 1980

* Brun-Cottan (1976) identified two distributions one <4–5 the other >4–5 similar to those reported by McCave (1975) and Harris (1977). Tailing was noted by Baker et al. (1979) below 3 µm. † Determined by scanning electron microscopy (S.E.M.).

SOME SAMPLING DIFFICULTIES Some of the many difficulties are given in the following four points. (1) It is necessary to measure the size distribution as soon as possible after sampling because the particulate mass, size, and composition can change on standing (Manuels in Eisma & Gieskes, 1977; Price, pers. comm.). (2) On plotting particle volume (mm3·1−1) against total suspended mass (mg·1−1) the gradient is equivalent to the sample density (Eisma & Gieskes, 1977). The points should fall between the gradients=1 (organic matter) and 2.65 g·cm−3 (taken as the average of inorganic detrital particles). The scatter below 2.65 was minimal but several points were above 1 with the conclusion that some organic particles have an inflated structure in water giving anomalous Coulter data. (3) The variation in total suspended matter and plankton species with season and location must be taken into account (see Eisma & Gieskes, 1977; Eisma & Kalf, 1979; data from the Southern Bight). (4) Plankton patchiness and the variation in size distribution with depth create particular problems. Results from the Atlantic are illustrated in Figures 9 and 10 (Pugh, 1978). The results serve as a cautionary note on how arbitrary the extrapolation of single station results to ‘world ocean’ models can be.

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634

Fig. 9.—Example of horizontal size distribution as determined with a HIAC cell (Pugh, 1978): the changes in the particle size spectra, chlorophyll a fluorescence, temperature, and nitrate concentration in the surface water between 20°50′ N:17°40′ W and 22°30′ N:17°40′ W.

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Fig. 10.—Example of vertical profiling at 21°05′ N:17°59′W using a HIAC cell (Pugh, 1978): individual particle counts m1−1 for five size ranges during the ‘down’ run.

SUMMARY There is no one practical technique available to cover the full size range of particles in sea water. As a rule of thumb, the range may be covered by: S.E.M.(+E.D.X-R.): 0.1 to 50 µm—samples on filter membranes; Impedance (Coulter): 1.0 to 40 µm—0.5 ml to 10 ml from water bottles; light (HIAC): 10 to 600 µm—1 l·min−1 in situ; and photographic (WHOI): >500 µm—in situ illuminated ‘block’. Distributions in the range 0.1–10 µm follow a power-law relation of segmented slope. Larger particle distributions have yet to be determined; the distribution over the full range may be bimodal. The conversion of number or volume distributions to mass distribution are inaccurate because of an inability to determine accurately individual particle density. S.E.M.-E.D.XR. analysis may help in this respect. As always, the limitations of season and selection of sampling site must be borne in mind in the subsequent treatment of data.

THE ESTIMATION OF FLUXES Lerman (1979) defines a flux as mass, energy, volume of particles equal to some proportionality factor multiplied by a driving force (advection, dispersal, diffusion). The driving force may be eddy diffusion (K), (vertical: 10−1 to 101 cm2·s−1; horizontal: 106 to 1010 cm2·s−1); molecular diffusion (D), (10−7 to 10−4 cm2·s−1); advective velocity (U) (surface waters: <10° to 102cm·s−1); particle settling; <10−3 to 10–1 cm·s−1; upwelling

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636

<10−6 to 10−5 cm·s−1; sedimentation <10−11 to 10−8 cm·s−1). Eddy diffusion arises from the dissipation of kinetic energy in the system and operates on concentration by decreasing concentration gradients. Advection is flow of material. Particle fluxes may be measured directly by sediment traps. Alterna-tively, they may be determined indirectly by the estimation of physical size, volume, mass distributions or by the measurement of radionuclide concentrations. SEDIMENT TRAPS Reynolds, Wiseman & Gardner (1980) recently published a review and annotated bibliography of all trap experiments covering the last hundred years. Over the past ten years increased use has been made of traps in the oceans, mostly on moorings and occasionally free floating (see Table IV). A correctly designed trap will give the flux (F) of material as M·L−2· T−1, where M, T and L are in units of mass, length, and time, respectively. The aim is to collect material of mass and composition representative of the flux. Finding a trap shape that will give a representative sample under all current velocity regimes is, as yet, an unresolved problem. In moving water, particles are carried into the trap by eddies where fluid exchange takes place. A particle will only be captured when it migrates from the turbulent water to the stagnant water at the base of the trap, therefore, the physical shape of the trap is of primary importance and the current flow is of secondary importance. Design of traps As can be seen from Table IV traps of various shapes and sizes have been deployed with various refinements. The effect of design on efficiency is discussed below. Flushing of traps on ascent during recovery can be prevented by use of shutters, either triggered by messenger or time-release (e.g. Rowe & Gardner, 1979). A further innovation is the use of time-sequencing traps (Kimmel, Axler & Goldman, 1977; Zeitschel, Diekmann & Uhlmann, 1978; Jannasch, Zafiriou & Farrington, 1980) to measure temporal variation in fluxes. Seasonal dependence is shown by the work of Deuser, Ross & Anderson (1981) who, over a two-year period of continuous trap collection (two month deployments) in the Sargasso Sea, found a variation of a factor of 3 in the flux (Table IV). Trapping efficiency Gardner (1980a,b) has attempted to assess the collection efficiencies of various designs in both laboratory flume experiments and in the field. The designs tested include domes, funnels with and without baffles, cones, cylinders, segmented boxes, plates, and narrownecked wide-bodied traps. The variation in trapping efficiency, both under- and overtrapping is considerable. Funnels fitted with baffles of cell aspect ratio 2 to 5 gave good results but a cylinder of aspect ratio 2 to 3 gave the most representative result in terms of measured sediment deposition in the flume. This did, however, depend on the velocity regime; an increase in current above 15cm·s−1 increased trapping with cylinders but decreased that of funnels. Hargrave & Burns (1979) experimented with various designs

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and found cylinders with an aspect ratio of 20 (or baffled cylinders of cylinder height to baffle cell ratio of 50) to have efficiencies of 95 to 100% in situ for currents of up to 8 cm·s−1. In long narrow traps a turbulence-free boundary layer is thought to exist that

TABLE IV Flux measurements from sediment traps

Shape

Trap Refinements collecting area

Place

Location Duration of Water Depth deployment depth off (m (days) (m) bottom (m) 63

2150

116

2.1 (May, 1969)

580

430

Box

1m2

16 compartments shutters Recovery with D.S.R.V.ALVIN

Tongue of the Ocean, Bahamas

Funnel

615 cm2

Shutters (Paired?)

Santa Barbara Basin

36°14′ N 120°01′ W

30 3.5 (Dec., 1969)

10

35 (Oct., 1971)

480 10

34 (Oct., 1973)

430 10

San Pedro Basin

33°30′ N

26.7 (May, 1972)

890

118°22′ W Soledad Basin

25°15′ N 112°41′ W

740

10 2.1 (May, 1969)

520

370 350 40 10

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638

24.6 (Oct., 1973)

370 10

Funnel

1.5m2

Paired

Sargasso Sea

Aspect, 8.4

Monterey 36°42′ N Bay 112°13′ W

Baffled Up to 8 per depth Free floating line

75

5581

214

1000

950

55°01′ W (or 55° 03′ W)

Baffled

Cylindrical 43cm2

31°32′ N

North Pacific Gyre

19 (nonupwelling) and 21 (up welling)

32°47′ N

7

750 300 5000

144°26′ W

4925 4425 3950

Shape

Trap Refinements collecting area

Cylindrical 177cm2

Place

Location Duration of Water Depth deployment depth off (m (days) (m) bottom (m)

Paired

Lower Cook

59°–60° 5–6 (MayN Aug., 1978)

Aspect, 2.7

Inlet, Alaska

151°– 154° W

Shutters

33–80

10

190

5288 4899

69.

4300

49.

1533

46.

(10 moorings)

Baffled Funnel

1.5

m2

13°30′ N

Paired

Tropical Atlantic

Baffled

(Barbados) 54°00′ W

Shutter

98

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165

Receiving cup,

Timer

North Central

15°21′ N

Pacific

151°28′ W

Sargasso Sea

31°33′ N

61

220

47.

5792 5414

11.

4814

110

3514

17.

1512

16.

210

11.

5581 4605

19.

1887

18.

375

13.

100

123

20

110

3520

50

205

1345

50

280

55°55′ W

Box

1m2

Baffled

Galapogos Rise

0°36′ N

234

2760

86°06′ W Cylindrical 314cm2

Western North

38°23′ N

Respirometers Atlantic

69°45′ W

Aspect 2.5

33°30′ N

Shutter

30?

76°15′ W 2 traps

2 traps

Shape

27°42′ N

675 100

78°54′ W

50

27°42′ N

645 100

78°54′ W

50

220

Trap Refinements Place Location Duration of Water Depth Ma collecting deployment depth off (mg·m area (days) (m) bottom (m)

Cylindrical 491 cm2

Aspect 3

Western 38°28′ N

10.1

2815

26

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640

North Shutter

Atlantic 72°01′ W

21 15

38°28′ N

10.7

2816

72°02′ W

500 100 13

38°19′ N

15.8

3577

69°37′ W 38°50′ N

518 118

5.8

2192

72°31′ W

36 33 30

Funnel

1.54m2

Baffled? Shutter?

Atlantic S.E. of Bermuda

62 (6 Apr. 1978)

All

All

62

4200

1000

62 55 64 68 60 52 61 61 63 52 (11 Aug. 1980)

17.5

prevents resuspension and flushing out of trapped material. This is dependent on the velocity scale, σb, associated with water motion at the bottom of the trap induced by flow across the top, thus for long cylinders, . Besides the aspect ratio, σb is dependent on the Reynold’s number associated with the flow over the cylinder mouth. The greatly exaggerated trapping of narrow-necked wide-bodied traps (and undertrapping of cones) is simply explained by Hargrave & Burns (1979). If the area of the mouth of the trap, Am, is much smaller than its base, Ab, and the suspended concentrations of particles are C0 in the water mass and Ct in the trap, then the number of particles with a settling velocity Us: (1) falling into the trap=UsC0Am·s−1

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(2) settling to the trap bottom=UsCtAb·s−1 thus, or . In turbulent flow, some of the water inside the trap is continuously being replaced by water of greater particle concentration, which leads to high flux estimates for The opposite is true of funnels where thus giving an under-estimate of the flux. Zeitschel et al. (1978) found considerable differences in collection between moored and free-floating traps. This effect was studied with dye experiments in Buzzards Bay by Staresinic, Rowe, Shaughnessy & Williams (1978). With moored traps, where currents were in excess of 15 cm·s−1, dye tracers developed into a turbulent wake over the trap whereas turbulence was not noted when dye was released away from the trap. This alteration in flow pattern may produce an uncharacteristic deposition. Such turbulent effects were not evident with free-floating traps. The logistics of free-floating trap recovery, however, may limit their use to areas of high sedimentation, i.e., short deployment. Two further ways to resolve the uncertainties of trap collection efficiencies have been undertaken; one by the comparison of the elemental (as well as radionuclide) composition of trap material with that of the underlying surface sediment, the other by intercalibration of various trap designs. In the southern Californian water basins Bruland, Franks, Landing & Soutar (1981) found the ratio of the collection rate to sediment accumulation rate to be 0.93±0.2 by use of stainless cones (baffled, 2500 cm2 collecting area, 120 cm long with 6 cm diameter collecting tube) on two-months deployment. Comparison of trap fluxes with 210Pb sediment accumulation rates in the Cook Inlet gave good agreement over a series of deployments (Larrance, Chester & Milburn, 1979). Knauer, Martin & Bruland (1979) used 210Pb as a guide to trapping efficiency and it varied by a factor of 10 from the estimated atmospheric supply rate. Large variations compared with sedimentation rates have been reported in other instances (flux for upper traps of strings 0.22 to 1.28 greater than sediments, deep traps 0.33 to 7.7 greater) (Soutar et al., (1977). The second experiment was done by Dymond et al. (1981) with paired cone, folding box, cylinder, and closing box designs. The trapped material and surface sediment major element composition were in good agreement, yet the measured fluxes during a 45-day deployment were between 370 and 774 g·m−2·yr−1 compared with a 25-yr record of 920g·m−2·yr−1. The factor of two difference between the designs is good but the discrepancy with the long-term record could be better especially as the experiment was conducted during a period of high run-off. The authors suggest the higher sediment record flux is due to a near-bottom input of detrital material. In situ sample preservation The decomposition of material over a long deployment (up to a year) may be retarded by the use of chloroform (Zeitschel et al., 1978), sodium azide (Honjo, 1978; Larrance et al., 1979), formalin+NaCl (ρ=1.07) (Knauer et al., 1979; Knauer & Martin, 1981) or possibly antibiotics. Preservatives may be added as solution or in a pelleted slow-release form. Zooplankton may also enter traps to feed, although this can be prevented in part by use of the above chemicals. Dead animals are generally hand-picked from the samples (Knauer

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642

et al., 1979). Three potential sources of error are the dissolution of trapped components when density gradients are used, the contamination of samples by additives, and reducing environments occurring in the traps during deployment. Summary of main points Excepting the intercalibration studies, Table IV summarizes the work on traps and the main points arising from the discussion are as follows: (1) Cylinders of aspect ratio 3 to 20 should be used (Hargrave & Burns, 1979; Gardner, 1980a), (2) the use of baffles of cell diameter to cell depth 1:4 is recommended in that they reduce the possibility of scouring by eddies and deter large animals from entering the trap to feed (Gardner, 1980b), (3) shutters, closed by timers or messenger, should be fitted to prevent sample loss on recovery, (4) poisons should be added to sample cups to prevent degradation of the samples, but animals killed on entering the trap must be removed before sample treatment, (5) traps offer a better sampling option than water bottles or continuous centrifugation, although their deployment and recovery is logistically more difficult and expensive. IN SITU LARGE VOLUME FILTRATION SYSTEMS The use of in situ large volume filtration systems (L.V.F.) has advantages over trap moorings in that many samples can be taken in a short period. The instrument can also act as a vehicle for a number of real-time monitors that provide essential physical data required for flux calculations. The main short-coming is the seasonal (even diurnal) variation of material settling through the column, particularly in shallow waters (Soutar et al. 1977; Bishop et al., 1980; Deuser et al., 1981). Most filtration systems operate only for hours, therefore, giving “instantaneous” samples on the time scale of biological cycles. The main features of various L.V.F. systems are given in Table V.Many workers have opted to supply d.c. power to the motor although Beer, Dauphin & Sholes (1974) and Bishop & Edmond (1976) used a.c. through large diameter conducting cables. The latter workers achieved volumes filtered of up to 32 m3 whereas the best to be expected from d.c. instruments is 1000 1 for 1-h operation (Gwilliam, Lawford & Simpson, in prep.). To some extent, the subsequent analysis of samples determines the type of filter to be employed, e.g., organics—glass fibre, S.E.M.—Nuclepore, but the limiting factor in filtration for a given system is the rate at which the filters clog and thus pore size and area of filters are important. Sheldon & Sutcliffe (1969) and Sheldon (1972) noted that heavily loaded filters retain particles

TABLE V In situ large volume filtration systems

Flow rates

Pore size Filters stack Diameter Power supply, Working

Place

Par

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(l·h−1

(µm)

arrangement

(cm)

169

extra monitors, special features

depth (m)

co ra (µg·

900→120

0.45 or Single plankton netting or larger sizes

20.5

115 Vac 19mm 3 conductor cable, brorize water-meter

Euphotic zone

180→40

0.45

Two in parallel

14.2

Pb-acid battery 6V, 93 A-h, nutating disc water-meter, depth-time recorder

5700

420→15

0.45 (Millipore HA)

Two in parallel

14.2

2.4 A at 115 Vac 4700 or Pb-acid battery 6V, 150 A·h, nephelometer, stainless steel turbine flowmeter, temperature, depth sensors, pinger, and Niskin water samplers

N. and 40–8 equatorial Pacific, 5 stations

30.5

480 Vac, 2–3 400 kw, 3 phase 3 hp pump, cast iron impeller Flowmeter

02°47′ N 16–5 08°51′ W, 6 depths

18000→1800 1.25, (glass Single fibre ×2)

Gulf of Maine, 51 stations, 166 samples

Cape Basin,

80–1 400

12–1

5 stations 18000→1800 ”

180?→30

Four independent

30.5

1.2 Single (units ? (Millipore used in membrane) strings of 6)

480 Vac, 2–3 1500 kw, 3 phase 3 hp pump, bronze impeller, valve actuator

00°40′ N 10–1 86° W, 3 casts

Pb-acid battery, 5500 pressure activated switch,

N. and S. – Pacific, 7 stations

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644

disc flow-meter 2000→200

1.0 Four (Nuclepore) independent 10 &60 (Estal) 100 (Nytal)

30

Solid Pb-acid 6000 battery packs transmissometer, particle size counter, conductivity flow-meters, temperature, depth, pinger, GO-FLO water samplers

Trials in – Loch Etive, Scotland and N. Atlantic

much smaller than the stated pore size. Flow can be assisted by the use of glass fibre filters or the stacking of filters of different pore sizes. Bishop & Edmond (1976) determined the retention efficiencies of Mead 935-BJ and Whatman GFF glass fibre filters. Both effectively collected >1.25 µm but two layers of Mead filters retained particles >0.8 µm at 95 % efficiency. Tamil and Poly filters are quantitative compared with 5-µm Millipore filters and 60 % efficient compared with 0.45-µm Millipore filters. For Nuclepore filters the size classification is based on 50 % removal efficiency of particles greater than the stated pore size e.g. 0.4-µm Nuclepore may retain 1-µm particles with 98 % efficiency (Bishop & Edmond, 1976). A detailed discussion of cap formation in Nuclepore filters is presented by Fan, Leaseburge, Hyun & Gentry (1978). Nuclepore filters are, however, better than glass fibre filters for subsequent S.E.M. work because the particles are retained at the surface. Also they retain less sodium and magnesium hence require less washing and give better gravimetric results (Cranston & Buckley, 1972). Filters should always be leached, washed and dried prior to use (e.g. Eaton, Likens & Bormann, 1969). Secondary filtration onto Selas Flotronics silver filters is recommended by Peterson (1976) for X-ray diffraction studies. Filters themselves, therefore, cannot easily be used for particle size differentiation and the methods described on pages 128–131 should be employed. A detailed knowledge of the size distribution is essential for flux calculations using L.V.F. techniques, because the mass flux is dependent on the diameter d of the particles. The flux, Fd, is calculated by assuming Stokesian settling:

(6) where [md]=the mass concentration of particles of diameter d (g·cm−3), USd=the Stokes settling velocity of particles diameter d (cm·s−1). The settling velocity is given by:

(7)

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where g=acceleration due to gravity at a given latitude (cm·s−2). p =in situ density of the particulate material (g·cm−3). p p density of sea water calculated from C.T.D. data (g·cm−3). s= d=the equivalent spherical diameter from the counter (cm). η=the viscosity of sea water, calculated from C.T.D. data (g·cm−1·s−1). More accurate estimation of Us is possible if particles of a given shape predominate where the dimensions may be determined, e.g., for faecal pellets (Bishop, 1977; Komar, Morse, Small & Fowler, 1981). The instrument under trial at the Institute of Oceanographic Sciences (U.K.) (Gwilliam, Lawford & Simpson, in prep.) is designed to give real-time measurement and calibration of ρs, T, P and size distribution followed by calculation of [md], Us and Fd using an assumed ρp (as used by Bishop, 1977) as a first estimate. Sampling points in the water column are selected on the basis of transmissometer profiles on the outward cast. The system outlined above will also carry four GO-FLO water bottles to collect water samples on commencement of filtration for the estimation of other related processes. Better estimates of [md] and ρp may be obtained after gravimetric analysis and S.E.M.E.D.X-R. analysis of samples as applied by Lambert et al. (1981). The mass flux for a given particle diameter, d, was given by

(8) which on integration produced the mass flux for the total log-normal distribution

(9) (see the original paper for the solution). For aluminosilicates corrections were made for the percentage of non-spheroidal particles, both to the mass and settling velocity terms. An in situ density of 1.6 g·cm−3 was assumed (Gardner et al., 1976) as was a 20% composition of aluminosilicates incorporated in organic aggregates. The resulting flux was 10 to 30 µg·cm−2·yr−1, between one and two orders of magnitude less than that observed in the deep ocean. Again the importance of large particle settling is stressed in order to explain such discrepancies. Summary In situ L.V.F. has one main draw back in that there is no direct access to flux as with sediment traps. Good estimates of fluxes require good estimates of in situ particle density and diameter. Seasonal variation may be assessed by repeated visits to sampling sites. Advantages over trap sampling are: (1) many samples from many depths and locations can be taken during a short time, (2) simultaneous measurement of the physical properties of the water column and particles can be made and (3) bacterial modification of the

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646

sample and biological and chemical contamination are minimal.

PARTICLE DYNAMICS In the following discussion some of the processes that affect particle size distribution and settling from surface to sediment will be examined. The mathematics of the geochemical processes are given in a recent book by Lerman (1979); many of the derivations of the equations quoted below are to be found there as are alternative approaches to the problems. For the purpose of this discussion the number distribution is assumed to approximate to the power-law

(10) for particles >1.0 µm. SETTLING OF PARTICLES AND DEVIATION FROM STOKES’ LAW The settling of a particle is assumed to follow Stokes’ Law which for the settling of a sphere is given by

(11) where

(12) and

(13) the equivalent sphere radius for a particle of volume vs. We have seen from

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Fig. 11.—Mean settling velocities of biogenic particles: from data in Smayda (1967) and Berger (1976) presented in Lerman (1979).

the preceding sections that participate matter consists mostly of biogenic material with significant quantities of terrigenic and authigenic minerals. Aggregates of mixed composition are also common. Due to the vast diversity of shape and density, virtually every particle will deviate from Stokes’ settling. Figure 11 illustrates this point for a variety of biogenic particles for the size range 1 to 500 µm. Most of the settling velocities fall between r1.2 and r1.7 (Lerman, 1979). The general definition of Stokes’ Law takes into account shape by the inclusion of the

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648

shape factor α:

(14) The decreasing order of Us with shape is ellipsoid prolate > sphere > ellipsoid oblate > tetrahedron and cylinder>needle> hollow sphere > ring > cap > disc which produces a full range variation in settling of about a factor of 2 for the same sphere radius (particle volume equals the equivalent sphere volume). The effect is that the equivalent sphere radius over-estimates settling. An aid to the reduction in errors is the estimation of particle density by S.E.M.-E.D.X-R. analysis and the calculation of viscosity for given depths based on hydrographic data because viscosity varies by a factor of 2 from warm surface water to deep cold water. Chase (1979) reported that the presence of dissolved organic matter in sea water and organic coatings on particles increased the settling velocity of particles in the range 5– 500 µm (based on the mean of two characteristic chords). The greatest effect was observed for small particles where variation of about an order of magnitude was noted, reducing to a factor of 2 for large particles. It is interesting to note that Brun-Cottan (1976) stated that Stokes’ settling does not apply to particles <5 µm because small particle electrostatic bonding, hydration and sorption of organic compounds reduce settling. Perhaps the Chase effect and the non-Stokesian behaviour suggested by BrunCottan (1976) in some way compensate for each other. We shall return to the importance of organic coatings in the discussion of ‘impact’ scavenging and adsorption-dissolution effects. At the other end of the spectrum, coarse particles are thought to supply most of the mass flux to sediments (Brun-Cottan, 1976) particularly in the form of faecal pellets and material. Bishop et al. (1977) estimated that faecal material supplied virtually all of the mass flux out of the mixed layer but were equivalent to only 4 % of the total particulate mass concentration. Similarly, in shelf water, the material collected in traps consisted entirely of faecal material (Soutar et al., 1977). Small, Fowler & Ünlü (1979) applied regression analysis to the settling velocity data on faecal pellets (Smayda, 1967, 1971; Fowler & Small, 1972; Wiebe, Boyd & Winget, 1976) and derived the relationship

(15) where V=the volume of the pellet and the correlation coefficient=0–92. Settling velocities ranged from 20 to 1000 m·day−1. In more recent experiments Komar, Morse, Small & Fowler (1981) found that copepod and euphausiid faecal pellets had settling velocities equivalent to a cylinder:

(16) where d and l are the cylinder diameter and length, respectively. They calculated ρp to be

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1.22 g·cm−3 (compared with a measured value of 1.23 g·cm−3). When faecal pellets were considered to be ellipsoidal the relationship with volume, V, is given by

(17) A plot of the right hand side against V produces a better fit than that given by Equation (15) because Equation (17) takes into account density differences between the sets of data. Thus, the 0.667 power dependence of V is better than the 0.513 dependence. The necessary correction (shape) factors for real samples are for the greater part unknown with the exception of faecal pellets. The use of laboratory produced faecal material for the measurement of Us, as for example, in Turner (1977) and Honjo & Roman (1978) has been criticized by Small et al. (1979) in that pellets produced under artificial conditions are less compact than those produced in situ and thus the true settling rate is underestimated. This may best be appreciated from the findings of Honjo (1978) and Honjo & Roman (1978); they found two distinct types of pellet, “green” that contained predominantly phytoplankton and “brown” that contained large quantities of non-biogenic minerals particularly clays. The former presumably were produced in the surface layers and escaped to deep water whereas the latter were excreted by mid- or deep-water dwelling coprophagic zooplankton. The non-biogenic minerals are either amplified by passing through the various trophic levels en route to the deep water or occur in deep water due to resuspension of sediment. The ∆ρ value of faecal pellets was assumed to be 0.1 to 0.5 (mean 0.2) by Bishop et al. (1977) based on data from Smayda (1971) and Fowler & Small (1972) and confirmed to be so by Komar et al. (1981). Faecal sedimentation offers a rapid means of transport to the ocean floor of small particles, such as diatom frustules, coccoliths, and clays, that would otherwise require up to decades to settle or, theoretically, totally dissolve in transit (e.g., Schrader, 1971; Honjo & Roman, 1978). Faecal material and other large particles are also subject to fragmentation in transit. Conversely, particles may be subject to aggregation. We shall first consider fragmentation. FRAGMENTATION A spherical parent particle of radius r, will break into (r/a)α fragments where a and a are positive constants and a is the smallest particle that will undergo fragmentation. If the parent population is given by dN/dr=Ar−b then the fragment number distribution is

(18) and the residual population

(19)

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where f is the fraction of unfragmented particles. The effect is to increase the slope of the number distribution since there are now greater numbers of small particles and the plot of log dN/dr against log r of the final combined population will deviate from a straight line. When fragmentation is complete, then f=0 and the fragment number distribution becomes

(20) where

(21) If the power exponents of the parent and residual populations are known, then α can be calculated. A special case exists for b=4 where it can be shown that the power term is independent of the multiplicity, α, and hence the residual and parent population plot as two parallel straight lines (log dN/dr against log r). The conservation of the gradient at 4 and decrease in the constant A with depth have been observed in the oceans (Lerman, Carder & Betzer, 1977; Lerman, 1979) but is not necessarily a universal feature (see Table III, p. 134). On dissolution of the pellicle membrane, faecal pellets may well fragment at depth becoming a source of fine particles in deep water (e.g., Honjo, 1980). Neither does it necessarily follow that the chemical composition (particle type) is constant. Again, S.E.M.-E.D.X-R. analysis proves invaluable in the examination of fragmentation patterns with depth. AGGREGATION Aggregation was said to be of little consequence in the general consideration of open ocean particle settling because the low particle concentration (<20 µg·1−1) produces a low probability of collision frequency and a lower frequency of successful collisions (Lerman, 1979). Nevertheless, with increasing emphasis on the elusive marine ‘snow’ (large ‘flakes’ of organic material) aggregation cannot be discounted (Honjo, pers. comm.). Lerman (1979) gives three mechanisms for aggregation: coagulation of small particles by Brownian motion, aggregation in velocity gradients as, for example, in the nepheloid layer, and scavenging by settling particles. The ‘trapping’ of sinking particles by platelets of organic marine snow would be of the last type. Lal (1980) put forward a scavenging model to explain what he considers to be anomalous results obtained for radionuclide estimates of settling velocity (Table VI and pp. 155–158). His argument runs that settling rates determined for 55Fe, 210Pb, 234Th, and 239Pu correspond to settling velocities of 5×10−4 to 10−3 cm·s−1, equivalent to particles of 6 to 10 µm in diameter, yet it is smaller particles that are in much greater abundance if the power-law number distribution holds. But the smaller particles have a much greater surface area, should, therefore, adsorb the radionuclides and thus have a settling velocity one order of

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magnitude less than that observed. A mechanism was suggested whereby the 1 µm particles (radius r0) adhered to the organic coatings of large particles (radius r) on impact; the rate (s−1) was given by

(22) The reciprocal of the integral (limit r0 to rmax) of the probability function corresponds to the mean time (T) between collisions; the effective velocity, impacting the particle radius r0 is:

, of the average particle

(23)

Values of T=0.25–1 yr and were the result with an increase of 50 % in the aggregated particle settling velocity. This explanation could well be superfluous if the observations of Chase (1979) are taken into account or nuclide adsorption is onto specific mineral particles of high density and well defined size as observed by Jedwab, (1979, 1980), Lambert et al. (1981) and Lane & Simpson (unpubl. data). The composition detail of the fine particulate matter (<10 µm) is illustrated in Figure 6 (p. 130). DISSOLUTION The rate of dissolution of particles, for example opal or calcite, may be approximated to a constant decrease in radius with respect to time of a spherical particle:

(24) When the settling velocity is expressed as

(25) then the rate of change of particle radius with depth from some arbitrary depth z=0 is

(26)

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After falling through some distance z, the radius, r, is given by the integration of Equation 26 to be

(27) where r0 is the radius at z=0. Total dissolution (r=0) will occur at a depth of

(28) The cube dependence of the dissolution depth means that large particles that settle rapidly through the water column to sediments will be little affected, whereas small particles (<10 µm) may dissolve within the normal transit time. One obvious effect of dissolution is to decrease the particle settling velocity and decrease the number of small particles in the spectrum with respect to depth. The mean settling velocity is expressed by the integral

(29)

(30)

where r and zr=0 are described by Equations 27 and 28. If the original particle spectrum at z=0 is given by is

then the change at some depth z due to dissolution

(31) The dissolution rate constant, ε, can also be derived by assuming a constant loss of particle mass, M, per unit area S:

(32) and the rate of mass loss per unit area

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(33) therefore

(34) A number of experiments have been carried out to examine dissolution rates by attaching sample cups containing the tests of organisms on moorings. Honjo & Erez (1978) gave the order in weight loss to be 23 to 36% for foraminiferans, 11.3 to 24% for coccoliths, and 12% for diatoms at depths >3600 m on 79-day deployment. Retardation of dissolution was probably due to the formation of organic coatings. Their model predicted that coccoliths and diatoms should be preserved in deep-sea sediments whereas foraminiferans and pteropods should completely dissolve. Thunell & Honjo (1981) proved the dissolution rate to be dependent on the position of the lysocline. Planktonic foraminiferans, during 123-day deployments showed weight losses of 5% at 665 m and 30% at 3791 m with fragmentation of whole animals increasing from 2% at 2869 m to 20% at 3769 m. The sequence of events leading to fragmentation and also the variability in dissolution between species was the subject of further study with benthic foraminiferans on a 61-day mooring (Corliss & Honjo, in press). Generally, pitting of the surface, the first evidence of dissolution, was noted at 978 to 2778 m and dissolution was extensive at 3978 m. At 5582 m and 5590 m there was drastic deterioration noted with most species in terms of breakage and erosion of the chambers. Work on coccolith dissolution in the Pacific has shown spatial variation dependent on kinetic dissolution rate, the degree of saturation of calcite in the deep sea, and standing crop in the productive layer (Goreau & Honjo, in press). The vertical transport of coccoliths is governed by the flux of faecal pellets from the surface; the pellets fragment at depth releasing coccoliths which then dissolve. The time variability of faecal supply and fragmentation and hence the time (seasonal) variation in coccolith dissolution was estimated. The combination of rapid settling and rapid dissolution produces a useful tracer for deep-sea cycling as particles with long residence times are horizontally advected over great distances along isopycnals which blur estimates of local short-term transformation rates. Takahashi & Honjo (1981) and Takahashi, Hurd, Asper & Honjo (1981) found that various species of radiolarians would take between 2 wk and 14 months to reach the sea bed (5 km) if no dissolution were to take place. The measured dissolution rates were sufficiently fast that the slower sinking animals would dissolve in the water column and those that reached the sediments would dissolve at the interface. FIRST ORDER SCAVENGING PROCESSES As Lal (1977) observed, the constancy of the particle size spectrum with depth in deep water indicates that the fragmentation process is dominant over dissolution (and aggregation) otherwise a radical change would become apparent at the lower end of the

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spectrum. The rapid decrease in particle number occurs below 1 µm (Lambert et al., 1981) whereas the lower limit for most spectra is below 1 µm. One means of maintaining distribution shape above 1 µm may be the formation of organic coatings which helps to explain the presence of small calcite particles below the lysocline (Lal & Lerman, 1975; Honjo & Erez, 1978). The decrease in particle number with depth, as indicated by a decrease in the constant A, can, however, be expressed by

(35) where k is the first order rate constant (Lerman, 1979). Thus, the particle concentration decreases with depth as a function of z and r. From this expression, the spectrum has a peak value

(36) when

(37) Thus, the maximum of particle concentration decreases as 1/zb/2 with depth or 1/z2 for b=4. If the mass concentration of material decreases with depth, but the shape of the distribution is unaffected, then the mean settling velocity remains the same. Thus particles are removed from solution in proportion to both number concentration and cross-sectional area synonymous to filtration:

(38) where ks is in cm−2·yr−1. The mass concentration varies as (39) i.e., the plot of dN/dr against log dr is dependent only on the z term. Lerman (1979) demonstrated that classic particulate mass concentration profiles can yield valuable information on transport and removal processes. In the example given, two segments were evident, both decreasing exponentially, the first from the surface (3

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µg·cm−3) to the base of the mixed layer (0.4 µg·cm−3), the second from the base of the mixed layer to the ocean floor (0.1 µg·cm−3). The mass-concentration profile can be expressed by

(40) where M0 is the surface concentration, M∞ is the steady state concentration at depth, and β is a constant (m−1). The half concentration depths (z1/2) for the two segments of the example quoted were 60 m for the surface layer and 1000 m for the deep water where

(41) The downward flux of material may now be given as

(42) where K is the mean vertical eddy diffusion coefficient and Ūs is the mean settling velocity of particles between radii r1 and r2. The first order removal of suspended matter is

(43) and the steady state diffusion-advection-reaction equation that results is:

(44) (negative k indicates removal). The differentials can be obtained by the differentiation of M as in Equation 40. M can be obtained from gravimetric data or calibrated optical methods. Thus, the constants in Equation 44 can be related by

(45) By measuring particle size distributions and estimating densities for various depths, Us can be calculated or else approximated by

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(46) The importance of knowing the large particle distribution is stressed because of the dependence on the largest particle radius, r2. An assumption that provides one solution to Equation 45 is that

but this must be used with caution. PARTICLE LAYERING

If upwelling of velocity, w, is introduced into the dissolution-settling velocity equations then the radius as a function of depth (Lal, 1977) becomes

(47) (which reduces to Equation 27 when w=0 and z1−z0=z). For small particles, a reversal of settling may occur when Us < w at a depth zr given by

(48) and dissolution will occur after travelling upwards a distance

(49) Lal (1977) demonstrated that the latter (zr−zr=0) lies between 10cm and 10 m for known values of ε. Calcite and radiolarian particles of <5 and of 20 µm, respectively, at z=0 could form such layers. Increased density or, more important, increased viscosity, both serve to decrease B in the Stokes’ equation (Equations 11 and 12). The vertical transport of particles may be written (Lerman, 1979) as:

(50) for steady state conditions where N is concentration (numbers of particles. cm−3) and chemical processes that effect particle concentration or size are given by R. Rearrangement of the derivatives in Equation 50 gives

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(51) Layers may occur where biological productivity is evident, e.g., plankton in surface water, and where the maximum mechanical stability arises i.e., where minimum eddy diffusivity occurs—the thermocline. From the pycnocline downwards K again increases. The gradient dK/dz is a velocity term and, therefore, a change in dK/dz affects the advective velocity of particles. Further discussion of diffusion-advection-reaction models are given below (pp. 159– 166). SUMMARY The assumption that particles follow Stokesian settling for solid spheres is an oversimplification. In order to obtain good estimates of settling velocities and hence fluxes more work on shape factors and accurate determination of in situ densitives of particles are required. Dissolved organic matter and organic coatings may affect settling rates. Faecal pellets are the best characterized particles in terms of settling and faecal matter are important as the major supply of material to sediments and of fine particles at depth via fragmentation. Estimates of fragmentation processes may be made from particle size spectra in conjunction with S.E.M.-E.D.X-R. analysis and bulk sample analysis. Experiments on calcareous and opaline tests have shown how dissolution also leads to fragmentation. Fine particles (<5 µm) occurring in the surface waters, whether by fragmentation, biogenic processes or from terrestrial sources may be horizontally advected along isopycnals and thus form layers. Layers may also be formed in upwelling areas. Simple scavenging may be estimated from vertical mass distribution profiles whereas work on scavenging and aggregation by ‘trapping’ on marine snow is just beginning.

RADIONUCLIDE GEOCHEMISTRY AND INFERRED PARTICLE DYNAMICS Radionuclides are of great value in the study of particle dynamics because of their differences in parent-daughter chemistries and their precisely known decay constants which, in favourable circumstances, can give the time dimension of processes. For the estimation of settling velocity it is required that the radionuclide be rapidly removed from solution onto a particulate phase. Radionuclides that undergo such scavenging were introduced to the oceans from atomic weapons tests in the 1950s and 1960s and also occur naturally in 238U (4n+2) and 232Th (4n) series nuclides. A full review of the work prior to 1971, including sources, distribution, processes, and chemistry, was published by the National Academy of Sciences (1971). Sackett (1978) presented a review of the literature up to 1977 with emphasis on the estimation of settling velocities (see Table VI). It is not intended to reiterate the early work in any detail here,

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but to concentrate on the most recent data. Two papers in particular will be discussed at length for they have brought together the ideas and findings of many researchers and produced models that both summarize and give new perspectives on radionuclide distribution. The first is on the distribution and behaviour of thorium isotopes (Bacon & Anderson, in press) and discusses reversible adsorption as well as the traditional assumption of irreversible uptake onto particles; the second examines two-dimensional models applied to the distribution of 210Pb (Spencer, Bacon & Brewer, 1981). RADIONUCLIDES FROM NUCLEAR DEVICE TESTING These radionuclides proved to be ideal tracers for the relatively simple estimation of settling velocity. Fallout from nuclear testing was introduced into the sea surface and the depth of penetration of the nuclide with time gave the settling rate of the phase with which the nuclide was associated. Broecker & Olson (1959) began the work with 14C and Somayajulu, Lal & Kusumgar (1969) were able to use excess 14C (over natural 14C levels) to estimate settling rates for calcium carbonate. The different rates at 2500 m and 3500 m of 2.2 and 1.6 m·day−1, respectively, perhaps reflected the effect of dissolution and fragmentation. Bowen & Sugihara (1965) noted the deeper penetration of 144Ce and 144Pm

TABLE VI Radionuclides: their use in the examination of oceanic processes and estimation of particle settling velocity: the settling velocities quoted correspond to particles of 2–10 µm in diameter of density 1.5 to 3g·cm−3; the half lives of other naturally occurring radionuclides are 238U 4·7×109 yr, 234U 2.44×10s yr, 228Ra 5·77 yr, 226Ra 1600 yr, and 222Rn 3·5 day

HalfNuclide life

Notes on production and geochemical cycling

Mean settling velocity Ocean and depth (m·day−1) Reference

(a) From bomb tests in the 1950s and 1960s 14C

55Fe

5,730 Must be corrected for naturally yr occurring levels; gives estimates of CaCO3 settling velocity.

2.6 yr Iron is rapidly removed from solution; may act as monitor for stable iron which scavenges other metal ions.

Deep-sea Pacific

1.6 & 2.1 Somayajulu et al., 1969

Deep-sea Pacific

2.2 Lal & Somayajulu, 1977

Deep-sea

1.0 Labeyrie et al., 1976

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239Pu

24 300 yr

239Pu is rapidly removed from solution.

185

Deep-sea Pacific

0.43 Lal & Somayajulu, 1977

Nearshore Pacific sediment data

≈1.0 Krishnaswami et al., 1973

Deep-sea Pacific

0.6 Miyake et al., 1970

Deep-sea Atlantic

0.4 Bowen et al., 1971

Deep-sea Pacific

0.29 Krishnaswami et al., 1976

Atlantic sediments

0.38 Noshkin & Bowen, 1973

(b) From natural decay series 234Th

24.1 day

All thorium isotopes are scavenged Mixed from solution; 232Th is the parent of layer 228Ra which decays to 228Th; the Indian short half life of 234Th makes it ideal for following shallow water and mixed layer processes; 232Th is introduced from the continents in minerals; 230Th paniculate concentration increases with depth with substantial supply from sedimentation. Nearshore Atlantic Mixed layer Deep-sea Atlantic

1.0 Bhat et al., 1969

10.0 Aller & Cochran, 1976 1.4 Matsumoto, 1975 0.33 Spencer et al., 1978 150 (large particles)

Nuclide Half-life

Notes on production and geochemical cycling

Mean settling velocity Ocean and depth (m·day−1)

Reference

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230Th

228Th

7.5×104yr

1.91 yr

660

Deep-sea Pacific

1.73 Krishnaswami et al., 1976

Whole water column Indian

1.73 Krishnaswami et al., 1981

All oceans Based on the intersection of Deep-sea 234Th 228-Th curves two values Atlantic for settling rates are obtained.

0.4 Broecker et al., 1973 0.36 Spencer et al., 1978 55 (large particles)

210pb

22.26 yr

210Pb is derived from 222Rn in Shallowwater deep water; in the mixed layer Pacific its main source is from atmospheric 222Rn which is degassed from the continents; the complicated 210Pb dynamics are discussed in the text.

0.4 Bruland et al., 1974

Shallowwater

0.1 Krishnaswami et al., 1975

Deep-sea Atlantic, Pacific

0.2 Craig et al., 1973

Mixed layer Atlantic

210Po

138.4 day 210Po is produced from 210Pb but has a shorter scavenging residence time; unlike 210Pb, 210Po is intensively recycled in the mixed layer.

0.05–0.5 Shannon et al., 1970

Deep-sea Atlantic

137 (large Spencer et al., particles) 1978

Deep-sea Atlantic, Pacific

1.2 Somayajulu & Craig, 1976

Mixed layer Atlantic

Mixed layer Deep-sea Atlantic

0.05–0.5 Shannon et al., 1970

0.3 Cherry et al., 1975 127 (large Spencer et al., particles) 1978

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with respect to 90Sr and 137Cs and suggested the association of the rare earth pair with a paniculate phase. Labeyrie, Livingstone & Bowen (1976) found a difference in the mean settling velocity estimates of 55Fe (350 m·yr−1) and 239Pu (195 m·yr−1) (Noskin & Bowen, 1973) and suggested that the two nuclides were adsorbed onto different particle phases. From Table VIa (based on Sackett, 1978) it can be seen that the settling rates decrease in the order 14C<55Fe<239Pu supporting the idea of varying particle speciation or adsorption kinetics. THE GEOCHEMISTRY OF THE NATURAL SERIES R A D I ONU C LID ES The two families that are most frequently applied to the elucidation of geochemical processes

are

the

238U(4n+2)

series

nuclides:

and the 232Th (4n) series nuclides: . Dashed arrows indicate one or more short-lived intermediates and the relevant half-lives are to be found in Table VI. For a naturally-occurring nuclide to be useful in the estimation of settling velocity, its half-life should ideally be within an order of magnitude of the time-scale of the process under investigation. These time scales may range from days for large particles settling in shallow water or the mixed layer to 100 years for small particles settling in the deep ocean. In a closed system on a sufficiently long time-scale, the rate of formation of a radionuclide becomes equal to its rate of decay, a condition termed secular equilibrium. In fact this condition is rare in the oceanic water column because of the different geochemistries of the elements in question. Uranium has its original source from land but its behaviour in sea water is dominated by its solution chemistry due to the formation of the stable [UO2(CO3)2]2− complex. Both 238U and 234U are uniformly distributed in the water column (Ku, Knauss & Mathieu, 1977). 232Th enters the oceans in minerals and hence is deposited to sediments. 234Th, on the other hand, is produced from the decay of dissolved 238U and is rapidly removed from solution. Its short half-life (24.1 days) makes it best suited for looking at fast processes. 230Th is also produced in situ and its particulate concentration increases with depth (e.g. Bacon & Anderson, in press). 228Th production is governed by the distribution of 228Ra; the behaviour of 228Ra and 226Ra is controlled by their solution chemistry. The main source of 226Ra is from the sediment from which it is mobilized after production from its thorium parent (Cochran, 1980). 210Pb is produced indirectly from 222Rn. 222Rn is degassed from the continental masses and transported over the oceans, thus 210Pb has an atmospheric source to the mixed layer. Also 210Pb has a deep-water source from the decay of 222Rn derived from 226Ra in the water and the sediment (Kadko, 1980). 210Pb is also rapidly scavenged by particles as is 210Pb, but the complex dynamics of these two will be discussed separately. Isotopes of thorium, lead, and polonium are present in the particulate phase at 4% to 20% of their total concentration, hence their removal by particles and the disequilibrium between parent and daughter pairs. Besides settling velocity, the disequilibria can be exploited to give information on eddy diffusion coefficients, upwelling velocity, scavenging rates, scavenging residence times and particle removal residence times.

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VERTICAL EDDY DIFFUSION AND UPWELLING VELOCITY Sarmiento et al. (1976) found the buoyancy flux, FB, to be given by

(52) where K is the eddy diffusion coefficient and is the Brunt-Väisälä stability frequency. FB is approximately constant at 4×10−6±4×10−6 for bottom water based on excess 222Rn profiles. Virtually the only source of 222Rn in the oceans is sediments, making it an ideal tracer. Values of K ranged from 5 to 440 cm2·s−1. The presence of horizontal motion, however, induces a further component dependent on the vertical gradient of horizontal velocity thus in the thermocline

rather than the ideal proportionality of

as in deep water (Lerman, 1979). The reason for the inverse relationship is obvious since the more stable a water column, the less susceptible it is to turbulent mixing. The stability frequency (or buoyancy gradient) may be calculated from hydrographic data:

(53) where σθ is potential density. Similarly, the linear segment of potential temperaturesalinity plots were used by Craig (1969, 1974) and Craig, Krishnaswami & Somayajulu (1973) to estimate scale distance or scale height:

(54) Li, Feely & Toggweiler (1980) obtained values for K and w based on 228Ra fluxes from sediments. Nozaki, Turekian & Von Damm (1980) determined the “apparent diffusion coefficient” in the thermocline to be 3 cm2·s−1 based on “D”=λ/µ2. Here A is the decay constant of 210Pb and µ is calculated from the relationship C=CM exp (–µz) where C is the excess 210Pb activity at depth and CM the 210Pb activity at the bottom of the mixed layer, height z above the bottom. The estimation of K and w using hydrographic and radionuclide measurements may prove useful in models that predict processes governing elemental and paniculate distributions in the oceans. SCAVENGING MODELS FOR THORIUM ISOTOPES The one-dimensional (vertical) diffusion-advection-reaction models for radionuclides were given by Craig (1969) and Craig et al. (1973) for dissolved species as:

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(55) and for paniculate species as:

(56) where Cd=the dissolved concentration of the nuclide, Cp=the particulate concentration of the nuclide, Kz=the vertical eddy diffusion coefficient, w=the vertical advective velocity, Us=the particle settling velocity, λ=the radioactive decay constant, Pd=the rate of production of the daughter from the dissolved parent, z=depth, t=time, and k1=the removal rate constant from solution to the particulate phase (adsorption rate coefficient) assuming irreversible uptake. Similar models for total concentration of nuclides were proposed by Tsunogai & Minagawa, (1978). The inadequacies of models used to elucidate thorium isotope distributions based on irreversible uptake have been recently discussed by Bacon & Andersen (in press). Their results and conclusions are described at length in the following. They and others (Krishnaswami et al., 1976) found that to explain particulate thorium profiles, 40% or more of the thorium isotopes must be incorporated into the particulate phase whereas, in reality, only 4% of 234Th, 15% of 228Th, and 17% of 230Th were associated with particulate matter. In an attempt to gain a more realistic model for deep-water thorium distributions Bacon & Andersen (in press) compared three simple processes: (a) irreversible uptake:

(b) irreversible uptake with fast particle removal:

(c) reversible exchange:

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where S is the net settling velocity, k−1 is the desorption rate coefficient and k2 is the rate coefficient of formation of large particles and scheme (a) is described by Equations 55 and 56. There are common assumptions in all three cases. (1) K and w may be ignored in that they are either of little significance where a short half-life is concerned against production, decay and removal rates, i.e., for 234Th, or insignificant compared to net particle settling velocity (also Spencer et al., 1978). (2) The system is at steady state (common to all previous work). (3) Pd is independent of depth, i.e. 238U and 234U exist only in solution and are evenly distributed throughout the water column. When these assumptions are applied to Equations 55 and 56 and the relevant rate terms are included for the three models then k and S values can be estimated for measured or known values of Cp, Cd, Pd and λ. Thus for models (a), (b), and (c), respectively, (i) the rates of uptake from solution onto particles are:

(57)

(58)

(59) Equation 57 gives a good approximation to steady state for 234Th because of its short half-life. (ii) The distribution between phases are, therefore, by rearrangement of Equations 57 to 59:

(60)

(61)

(62) In model (a) k1=0.42·yr−1 and in models (b) and (c) k1=0.52·yr−1 due to the additional loss terms; k2=2.6·yr−1. (iii) The dissolved concentration:

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(63) i.e., the dissolved concentration is constant with depth

(64) i.e., the dissolved concentration increases with depth due to the presence of the desorption term, (iv) Particulate thorium is given by:

(65)

(66)

(67) Spencer et al. (1978) also used Equation 65. Solutions to Equations 65, 66 and 67, for the upper boundary condition of Cp=0, z=0 (i.e., no transport across the air-sea interface) and after substitution of the relevant Cd values (Equations 63 and 64), were shown to be

(68)

(69)

(70) Because 230Th has a very long half-life, the decay terms are negligible and, therefore, a linear increase in Cp would be expected with depth. Equations 68 and 70 reduce to

(71)

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and hence S can be estimated from the gradient (see also Krishnaswami et al., 1976). The frequently used irreversible uptake model gives a good fit for 234Th data but not for 230Th. For the second and third models good fits to the data were found suggesting rapid particle transfer (see also Spencer et al., 1978) or extensive desorption. The reversible uptake model, however, also predicts the increase in dissolved concentration with depth found not only by Bacon & Anderson (in press) but also by Moore (1981) and Nozaki, Horibe & Tsubota (1981). Overall settling velocities of 100 to 1000 m·yr−1 were predicted and application of the model to the data of Nozaki et al. (1981) gave k1=1.5·yr−1, k−1=6.3·yr−1 and S=380 m·yr−1. In addition a fourth model incorporating all three rate constants was produced where:

(72) and for 230Th:

(73) and

(74)

From the previous models k1 was assumed to be 0.52·yr−1 and

.

0.1·yr−1,

The data required that k2 be ≤ therefore, desorption must dominate over settling for thorium and hence many adsorption-desorption cycles must occur before removal to sediments. Furthermore, the distribution of 234Th is in chemical disequilibrium because of the 234Th short half-life, therefore Equation 62 holds, whereas the 230Th distribution is in chemical equilibrium because of its long half-life. Thus . The measurement of both nuclides in the same sample allows for both the kinetics and equilibrium state to be defined: the adsorption rate will be dependent on particle concentration (surface area) and desorption rate independent of particle concentration. Bacon & Anderson (in press) stress that 230Th is in a state of dynamic equilibrium; this is of the utmost importance since the residence time with respect to removal for any element can be given by

where τe=the residence time of the element, τp=the residence time of the particulate matter, and [Ep] and [Et] are the particulate and total metal concentrations. The

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distribution ratio may either be measured directly or calculated from equilibrium constants (Balistrieri, Brewer & Murray, 1981). With equilibrium models for surface adsorption, Balistrieri et al. (1981) found that metal adsorption onto suspended particulates was most probably controlled by organic coatings on particles and that specific mineral phases played only a minor rôle in element uptake. Bacon & Anderson (in press) calculated τp to be 5 to 10 yr for the elements, Th, Mn, Pa, Pb, and Cu which have τe values ranging from ≈30 (Th) to 1000 (Cu) yr. The equivalent mean particle settling velocities are ≈1 to 3 m·day−1 with adsorption onto particles of a time averaged diameter 10 to 30 µm. These estimates emphasize the importance of fragmentation (e.g., of faecal pellets), aggregation (e.g., marine ‘snow’), and the presence of organic matter in the consideration of particle dynamics as discussed above (pp. 148–152). Reversible adsorption also helps to explain previous discrepancies in settling rates noted for different isotopes at the same station as calculated for 210Pb and 230Th (Krishnaswami, Sarin & Somayajulu, 1981), yet these authors neglected the boundary removal and horizontal redistribution which are more important for 210Pb than 230Th, thus 210Pb values for Pd and S would be too high. The effect of particle concentration on dynamic equilibrium is supported by the work of Li et al. (1980) and Moore, Bruland & Michel (1981) in coastal and high productivity areas (see also Table VI). The decrease, however, in thorium concentration from midwater to sediment that is sometimes observed (Bacon & Anderson, in press) is not taken into account by the above models and bottom boundary layer scavenging (of the type described below for 210Pb) must be invoked (Spencer et al. 1981). THE 210Po CYCLE AND TWO-DIMENSIONAL MODELS FOR 210Pb Bishop et al. (1977) found a greater enrichment of 210Po relative to 210Pb in the <1 µm particles consistent with a “nutrient-like” behaviour. Special attention will be given to 210Po because of apparent involvement in the biological cycle and to 210Pb because of its unusual behaviour in deep water. The short mean residence time of 210Po in surface waters of 0.6 yr (Bacon, Spencer & Brewer, 1976; Nozaki, Thomson & Turekian, 1976) was suggested by Cherry, Fowler, Beasley & Heyraud (1975) and Beasley et al. (1978) to be due to zooplankton faecal pellet production. Concentration factors of 104 for 210Po can be compared with 210Pb values of 102 (Heyraud & Cherry, 1979). These values partly tie in with estimates of surface water residence times obtained in various seas and oceans of 0.4 to 1.2 yr and 9 to 24 yr for 210Po and 210Pb, respectively, (Turekian, Kharkar & Thomson, 1974). A shorter mixed layer residence for 210Pb of 2.5 yr was calculated by Bacon et al. (1976). 210Po appears to be intensively recycled (about 50%) in the thermocline where there is a large excess of 210Po compared with 210Pb (Bacon et al., 1976; Thomson & Turekian, 1976) (cf. particle layering). Moore et al. (1981) had difficulty in explaining excess quantities of radioisotopes in trap samples collected in the Santa Barbara Basin based on in situ production and atmospheric fluxes; 210Pb and 210Po were, respectively seven and thirty times greater than expected. To account for these observations, it was suggested that either water passing into the area was stripped of nuclides by the high flux of particles and/or freshly

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deposited sediments were being resuspended from the basin edges and transported back to the basin on a time scale of a few years. Opposite boundary effects have been noted in the deep ocean where there is an exponential decrease in dissolved 210Pb in deep water down to the sediment interface, but with no corresponding increase in the 210Pb paniculate concentration (e.g., Goldberg, 1963; Bacon et al., 1976; Nozaki & Tsunogai, 1976; Somayajulu & Craig, 1976; Thomson & Turekian, 1976; Nozaki et al., 1980). Following a review of the data (Bacon, Spencer & Brewer, 1978), Spencer et al. (1981) produced a two-dimensional diffusion model that offered some explanation for the observed effects. At steady state, the distribution of 210Pb is a balance between input and output:

(75) where Kx is the horizontal eddy diffusion coefficient, F0 is the first-order removal by adsorption (yr−1) and F1e−µz is biological uptake decreasing exponentially with depth. The model used boundary fluxes of

(76) where (x, z) is the piston velocity describing the flux of 210Pb to the boundary; this varies from east to west and bottom to top (where—Kz equalled the atmospheric flux of 0.6 dpm·cm−2·yr−1). Worthy of note is that the double differentials were approximated to

(77) where x0=the horizontal grid spacing. The boundary values are calculated by using a point outside the boundary. The best fit to the data used values of Kz=1 cm2·s−1, Kx interior=5×106 cm2·s−1 and Kx boundary= 184.8 cm2·s−1 (based on the work of Okubo, 1971); the reader is referred to the original paper for other parameters. Figure 12a shows the original data and Figure 12b the fit of the model. The boundary effects have been reproduced with remarkable similarity i.e., high surface concentration, lowest activities at the western boundary, maximum concentration effect to the east at 4 to 4.5 km depth, the intrusion of the minima at ≈1 to 1.5 km and west and the decrease in activity in deep water are all apparent. Of the fluxes, just over half is accounted for by radioactive decay, a quarter by biological uptake and, most significantly, the boundary uptake is equivalent to the in situ adsorption. Large fluxes, in excess of the atmospheric supply due to large particle settling, conform to the boundary effects (0.6 dpm·cm−2·yr−1) and only low deepocean fluxes are required to give the observed deep water gradients (0.04 dpm·cm−2·yr−1). A second approach was also considered. Figure 12c was produced by assuming in situ

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adsorption and biological uptake as a function of x and z, i.e. there is no boundary uptake but higher adsorption occurs at the boundaries than in the interior. Thus

(78) To some extent, the model fits the known particle distribution and fluxes in the Atlantic. The increase in adsorption, and hence particle concentration, required from the interior to the western boundary is unreasonably high in

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Fig. 12.—a, dissolved 210Pb (dpm·100 kg−1) in west-east section (15°–30° N) in the Atlantic (after Bacon, 1977); b, model 210Pb distribution including side and bottom scavenging; c, model 210Pb distribution with no boundary scavenging and variable in situ adsorption in water; after Spencer, Bacon & Brewer (1981).

that a 40-fold increase in F0 must be invoked. These increases must also operate throughout the water column. To explain the deep-water boundary effects it was suggested that hydrous iron and manganese oxides act as scavengers, a suggestion also put forward by Thomson & Turekian (1976). Iron and manganese are remobilized after dissolution in anoxic sediments; such environments are encountered in areas underlying high surface productivity. On reaching oxidizing overlying waters, hydrous iron and manganese oxides co-precipitate 210Pb. There is evidence for such behaviour in the anoxic waters in the Cariaco Trench (Bacon et al., 1980). Flocculent iron particles are certainly a common feature in S.E.M.-E.D.X-R. analyses of suspended particulate material (Lane, Simpson & Nott, unpubl. data). Models of this type, using detailed chemical and physical data collected on a threedimensional grid basis, indicate a means towards understanding the relative significance of many processes and fluxes.

CONCLUDING REMARKS Sample collection must provide sufficient uncontaminated material for full chemical and morphological analysis which generally precludes water-bottle and centrifugation methods. Sediment traps and in situ large volume filtration systems can provide such samples but both have limitations. Traps can be logistically difficult to deploy and recover and there is a financial restraint that limits duplicity of moorings and hence sample numbers. Although precision with traps was shown to be within a factor of two for various designs, the accuracy of sample collection has yet to be proved. Traps do have the advantage of producing a long term record of sedimentation and fluxes. L.V.F. systems have the advantage of producing many samples at many stations with the facility for continuous monitoring of physical conditions and simultaneous collection of water samples. L.V.F., however, only produces ‘instantaneous’ samples and not long term records which may lead to errors in flux calculations. Fundamental to particle dynamics and fluxes is the knowledge of how particle distributions change with depth and how the mineral chemical and biological composition of the distributions also alters. In situ particle counters that cover large ranges in size, from the numerous small particles to the rare but important large particles, i.e., faecal pellets and marine ‘snow’, will be of great value. Central to analysis is the semi-quantitative data produced by S.E.M.-E.D.X-R. analysis. This facility provides information on fragmentation, aggregation, dissolution, adsorption, and biological species variation not directly observable by other methods.

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The importance of organics in sea water and particles has been suggested in connection with enhancement of settling velocity, retardation of dissolution, an aid to aggregation and control in adsorption, all of which are poorly understood and require investigation. Radionuclide work has produced a valuable insight into the complicated biological and geochemical processes. From the radionuclide studies it has been shown that the element with the shortest known oceanic residence time, thorium, is in dynamic equilibrium with respect to adsorption (from 230Th data). The implication with respect to stable elements is that they also are in dynamic equilibrium. Whereas radionuclide studies give the overall net result of particle and element dynamics, the detail of the processes for elements and nutrients must be resolved by chemical speciation models and measurements and include models for aggregationdissolution-fragmentation of the particles themselves. The past decade has been especially productive in terms of ingenuity, results, and ideas, and the future promises to be equally as stimulating.

ACKNOWLEDGEMENTS The author wishes to thank Dr R.J.Morris for his helpful criticism, Dr C. Lambert for her comments, and Dr J.Thomson for discussions on radionuclide geochemical cycling and processes. The ‘voluntary’ refereeing of the manuscript by Drs J.Bishop and M.Bacon is gratefully acknowledged.

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BIOLOGY AND ECOLOGY OF MARINE OLIGOCHAETA, A REVIEW OLAV GIERE Zoological Institute and Museum, University of Hamburg, Martin-Luther-KingPlatz 3, D-2000 Hamburg 13, Federal Republic of Germany and OLAF PFANNKUCHE Institute for Hydrobiology and Fishery Science, Hydrobiological Department, University of Hamburg, Zeiseweg 9, D-2000 Hamburg 50, Federal Republic of Germany

Oceanogr. Mar. Biol. Ann. Rev., 1982, 20, 173–308 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION Marine Oligochaeta, hardly mentioned in Stephenson’s (1930) classical monograph, have received increasing attention during the past 50 years. The number of known species has increased rapidly, especially with the growing evidence of the diverse oligochaete fauna in sublittoral biotopes (Erséus, 1980a); this leaves taxonomic considerations in continuous state of flux. Biologically and ecologically, marine oligochaetes have received growing attention, at least in some special aspects (e.g. meiofauna, Lassèrre, 1976; zoogeography, Dzwillo, 1966; Erséus & Lassèrre, 1976; Erséus, 1980a). The first detailed ecological study on marine and brackish-water oligochaetes came from Remane’s school: Knöller (1935a) investigated the oligochaete fauna of the German Baltic and North Sea shores. His work was later continued by Bülow (1957). These papers not only covered the typical upper littoral habitats, but focused also on the interesting forms living in the “Küstengrundwasser”. The comprehensive monograph of Backlund (1945) on wrack fauna contributed important details on the two most common shore oligochaetes, Lumbricillus lineatus and Enchytraeus albidus. Since then, more details have been accumulated, e.g. on the life history, autecology, distribution patterns, production, and response to pollutants. Nevertheless, even quite recent general reviews on oligochaetes (e.g. Stolte, 1969) still widely disregard marine oligochaetes and there exists no specific compilation on their ecology and biology. Even in the latest proceedings on Aquatic Oligochaete Biology (Brinkhurst & Cook, 1980) freshwater forms were given a predominant rôle. Therefore, it seemed appropriate to draw attention to the marine Oligochaeta by compiling the scattered data in this contribution.

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With its emphasis on the truly marine and brackish-water species this review will deal only marginally with the numerous euryhaline freshwater species which occasionally leave the borders of limnetic habitats. Morphological and taxonomical details will only be included in passing for general characterization of dominant groups (see p. 174 ff) and, if relevant, for ecological and biological aspects. Anatomy, sexual biology, and ontogeny will not be covered here, as details on marine species either do not differ much from descriptions in general reviews or have been dealt with elsewhere (Brinkhurst & Jamieson, 1971; Lassèrre, 1975a). Particularly in the anatomical pattern of their reproductive organs, marine oligochaetes have maintained almost identical structures, clearly homologous to those in their freshwater relatives. Even the truly marine Phallodrilinae, although possibly originating from ancestors already populating the marine biome, have not developed basically new arrangements. This consistency proves that despite their recent radiation marine Oligochaeta have descended from limnetic forms. Fully aware of the fact that ecological and biological data at present are far from rendering a definite picture and will continuously be supplemented and corrected, we want to form a basis which mirrors the ‘state of the art’ today (spring 1981) and facilitates quick information for non-specialists as well as specialists in this field. We will be satisfied if it can lead to a more correct assessment of the biological importance and ecological relevance of this group within the marine environment.

BRIEF CHARACTERIZATION OF MARINE OLIGOCHAETE GROUPS The more than 200 marine oligochaete species, described so far, belong mainly to two families, the Tubificidae and Enchytraeidae. Truly marine representatives of other groups (Naididae, Aeolosomatidae, Megascolecidae, and Phreodrilidae) are restricted to a few scattered species. Because ecological and biological data on Aeolosomatidae and Phreodrilidae are almost completely lacking, these groups will only be mentioned briefly, mainly in the following ‘diagnosis’. Prominent features in the taxonomy of oligochaete families are (a) position of genital organs and their accessory elements, and (b) of less importance—the structure and arrangement of the setae (Fig. 1). Development of genital structures in specific segments is a clearly discriminating and a fairly consistent feature on the family level although, in exceptional cases, multiplication of spermathecae, supernumerary gonads, unilateral development or fusion of genital organs, their shift to abnormal segments or complete absence have been described (Stephenson, 1930; Dumnicka & Kasprzak, 1979). These ‘deviations’ can be found regularly and characteristically in certain species (e.g. Grania monospermatheca Erséus & Lassèrre, 1976; Phallodrilus postspermathecus Erséus, 1979b) or even genera (Aktedtrilus, Coralliodrilus, Uniporodrilus, Inanidrilus, Jamiesoniella; Erséus, 1980a; Jolydrilus, Heterodrilus; Marcus, 1965) or as anomalies in singular specimens (e.g. Lumbricillus sp. with only one spermatheca; Grania sp. with paired spermatheca in segments V and VI; pers. obs.). Setation is generally more variable and does not, as a feature per se, allow reliable group separation. Considerable changes both in setal shape and arrangement (number per

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segment) can occur intraspecifically and also

Fig. 1.—Diagrammatic summary of taxonomical characteristics in oligochaete families with marine representatives: in non-shaded zones (generative segments): black portions, testes; circles, ovaries; other structures, invaginations, spermathecae, efferent ducts and related organs; (modified after Lassèrre, 1975a).

within populations or even individuals. There is some evidence that impact of rigorous environmental conditions (e.g. extremes of salinity, pollution, ionic composition) may influence setal formation (Hagen, 1954; Loden & Harman, 1980; Milbrink, 1980) as well as the number of segments (heavy wave action, Hagen, 1951; Stolte, 1969). FAMILY DIAGNOSIS (Fig. 1): Tubificidae: dorsal and ventral setae frequently dissimilar, mostly bifid crotches, hair setae or derivatives; often with specialized genital setae. Paired spermatheca (in segment

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X) close to genital segments, male pores in XI, chitinous penis sheaths frequent, male funnels in X. Paired atrium and prostate glands usually present. Generic classification based mainly on structure of genital organs. An ever increasing number of species is described from sublittoral habitats. These are mostly small (meiobenthic) whitish worms. The more conspicuous littoral species are relatively large (10–70 mm), often reddish in colour and live in muds or sands rich in decaying debris. Enchytraeidae: shape of dorsal and ventral setae usually identical, mostly pointed simple needles, straight or slightly sigmoid, no genital setae. Paired spermathecae in V, 5/6 segments anterior to genital segments, male pores in XII; with penial bulb, but penes, prostate glands and atria lacking. Usually with conspicuous septal glands in V–VII. In generic classification, somatic features play a dominant rôle which contrasts to species identification (Erséus, 1980b). Naididae: dorsal and ventral setae usually dissimilar; mostly bifid crotches with nodulus (ventrally), long hair setae (dorsally) and short needles of variable shape and number; often penial setae, if mature. Asexual division, originating from budding zones, prevailing. Spermathecae (mostly in V) in the segment which bears paired testes and male funnels. Male pores and atria in subsequent segment. No penes; prostate cells often diffuse. Small, often transparent species, mostly epibenthic and often actively swimming forms. The almost exclusively limnetic family has few euryhaline representatives. Aeolosomatidae: A group of uncertain systematic relationship to the Oligochaeta (Brinkhurst & Jamieson, 1971). With dorsal and ventral bundles of long hair setae, often also with slender crotches or needles. Small worms (few mm) with “colour glands” in their transparent epidermis, prostomium ventrally ciliated; brain an epithelial plexus in contact with epidermis. Asexual division characteristic for the family, hence, specimens with reproductive organs rare. Genital organs of Aeolosoma maritimum, the single marine species described so far, not known. Phreodrilidae: The few littoral species with two ventral crotches per bundle, dorsally with needles or hair setae only, no gizzard. Spermathecae in XIII, often extending more backwards, posterior to the pair of testes in XI; male funnel in XI, tubular prostate and male pore in XII; here also paired ovaries, female pore in XIII. Megascolecidae: The few littoral representatives of these large (100 mm in length) ‘earth worms’ have eight regular rows of single setae, slightly serrated at their tips. Gizzard rudimentary. Spermathecae in VIII and IX; male pores in XVIII, in conjunction with prostates; a pair of testes each in X and XI; paired ovaries in XIII, female pores in XIV.

METHODS FOR BIOLOGICAL AND ECOLOGICAL OLIGOCHAETE STUDIES SAMPLING OLIGOCHAETE MATERIAL Depending on the aim of the investigation, sampling stations for oligochaetes should, if

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possible, be spaced along profiles, in eulittoral shores usually perpendicular to the water line or, if a larger area is to be examined, along a regular sampling grid dense enough to allow statistical analysis of the results. For studies on the horizontal micro-zonation, use of a “spacing plate” for sampling (Steele, Munro & Giese, 1970) may be helpful in assessing in all stations geometrically identical random samples. Vertical microdistribution of smaller oligochaetes is often investigated with Perspex tubes of varying length and construction. Through a series of holes in the walls electrodes can be inserted for recordings of physiographical background data. Larger (macrobenthic) oligochaetes are usually taken by square metal corers preferably with one movable wall for opening (e.g. McLachlan, Dye & Ryst, 1979). In submersed bottoms, samples should be taken by large-diameter corers pushed into the sediment by hand and, in deeper water, by diving or with use of a single or multiple gravity sampler (Brinkhurst, Chua & Batoosigh, 1969; Hakala, 1971). The various grabs, commonly used in macrofaunal surveys (reviewed in McIntyre, 1969), proved to be less reliable because (a) their depth of penetration is too small to sample quantitatively the oligochaetes which often populate the layers deeper than 10 cm below the bottom surface (Pfannkuche, 1980a) and (b) they regularly give underestimates compared with diver-collected cores (Holopainen & Sarvala, 1975; Elmgren, 1976). EXTRACTION AND EVALUATION OF OLIGOCHAETE SAMPLES Biological and ecological results must usually be based on rich, statistically evaluable material. In rare cases only, the natural population density will be high and the body size of the worms large enough to allow their collection by simply picking them up from the substratum. This method is probably suitable for species such as Enchytraeus albidus, Lumbricillus lineatus and L. rivalis, amply found on wrack weed beds and also in Peloscolex benedeni and Tubifex costatus, abundant on most tidal flats. The smaller naidids, as well as all interstitial forms have to be extracted from the sediment although they are often extremely common in shallow brackish bights and in lenitic beaches, respectively. Reliable extraction results can be obtained by repeated gentle washing with stirring and subsequent decanting through sieves (Dybern, Ackefors & Elmgren, 1976). Sieving often involves two major sources of error: choice of inadequate mesh size(s) and mechanical damaging of the worms’ bodies, particularly breaking of zooids in the fragile naidids. The importance of using relatively fine sieves even for larger oligochaetes is shown by the authors’ experiments in which from a mixture of preserved Peloscolex benedeni, Tubifex costatus and T. pseudogaster only 26% were retained in the 1000 µm-sieve, but 49% in the 500 µm-sieve and still 24.5% in the 200 µm-sieve. A more convenient separation of (fixed) fauna from mud or fine sand samples is the ‘flotation technique’ with sugar or Ludox (Du Pont de Nemours) solutions as separating media (Heip, Smol & Hautekiet, 1974; De Jonge & Bouwman, 1977; Nichols, 1979). Combined with centrifugation, this method represents an effective and convenient procedure for oligochaetes as tested specifically by Kajak, Dusoge & Prejs (1968) and in own studies. In many mixed sand-mud samples, previous ‘elutriation’ (see over) considerably diminishes the sediment volume containing the animals and thus facilities

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flotation and reduces Ludox consumption. The best way of separating meiobenthic oligochaetes from medium to coarse sand is ‘elutriation’, i.e. rinsing the sample by a water jet from below and making use of the difference in settling velocity between animals and sand. The widest applied modern elutriation method is that modified by McIntyre for meiofauna extraction which proved to be superior in efficiency and effort (time) to most other procedures (Uhlig, Thiel & Gray, 1973). A more sophisticated elutriation apparatus adapted to the separation of animals even from fine sand has been developed by Tiemann & Betz (1979). Also based on the elutriation principle is the complicated “Oostenbrink apparatus” (Oostenbrink, 1960), somewhat confusingly called a flotation device in a recent modification by Fricke (1979). Advantage of elutriation is its application both to live and preserved animals, its shortcomings are the limitation to small interstitial fauna. The ‘sea water-ice extraction’ method (Uhlig, 1964), commonly used for ‘soft meiofauna’, has often been proved unreliable with interstitial oligochaetes (Uhlig et al., 1973) which tend to remain in considerable numbers in the sediment core. Hence, data on (mesopsammic) oligochaetes obtained by this method should be interpreted with caution. In distribution studies, arrangement of sampling stations and amount of sample units should enable statistical analysis in order to cope with the problem of contagious occurrence (Abrahamsen & Strand, 1970) typical for many marine oligochaetes (Giere, 1971d). Replicate sampling with tolls rendering representative numbers of worms and appropriate statistical treatment which considers over-dispersion (Abrahamsen & Strand, 1970), are necessary for relevant conclusions on oligochaete distribution. Abrahamsen (1969) states as a rough approximation about 40 specimens per sample unit are necessary for statistical reliability. OBTAINING BIOMETRICAL DATA FOR ECOLOGICAL ANALYSIS Determination of body length and width measurements should be based on relaxed, live worms, preferably anaesthetized (7% magnesium chloride; 1% xylocaine or lidocaine, MS 222 Sandoz; 0.5% propylenphenoxetole). When there is only fixed material, the specific shrinking factor has to be determined (Wiederholm & Eriksson, 1977). Determination of surface area based on mathematical calculation for vermiform (=cylindrical) bodies (Andrássy, 1956; Abrahamsen, 1973) is not only very tedious, but volume data derived from these methods proved to be markedly higher than those assessed by water displacement in small graduated glass tubes (Giere, 1975); the latter method apparently, takes into account the uneven, segmented body of oligochaetes. A detailed description of procedures for assessment of body volume in oligochaetes is given by Lassèrre (1976). The problem mentioned above also limits the value of oligochaete biomass estimations on the basis of length-volume regressions and density determinations. Growth calculation from length-weight regressions (O’Connor, 1963; Edwards, 1967) may also be restricted and, thus, led Standen (1973) to use the number of segments for measurement of growth. The development of ultrabalances with an accuracy of 1 µg has made the determination of dry weights of single worms possible; biomass and growth calculation should, therefore, be based on weight. Measurement of wet weight, however, be it from

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live or preserved material, is known to be unsatisfactory due to evaporative losses during weighing; these are considerable even when batches of worms are taken. Lundkvist (1978) tried to avoid desiccation problems by inserting the worms into a pre-weighed drop of ‘silica fluid’ by which she obtained fairly constant values even using single specimens. Significant changes in both wet and dry weight as compared with ‘fresh weight’ occur after preservation. This phenomenon, which should be taken into account in biomass calculations from preserved oligochaetes, was studied in detail by Wiederholm & Eriksson (1977) who found an average fresh weight-loss of 13% (wet) and 89% (dry), respectively, for several common alcohol-preserved (limnetic) tubificid species. On the other hand, Williams, Solbé & Edwards (1969) could not find significant differences between fresh weight and preserved (wet) weight of Lumbricillus rivalis and Enchytraeus coronatus. They used, therefore, preserved material for a very well correlated length-weight regression line. In any case, higher precision and, hence, better comparability can be achieved expressing weight as dry weight. ‘Normal’ drying of clean, defaecated worms (preferably 24 h at 80 °C) eliminates any water from the material; subsequent combustion (1 h at 500 °C, see Hallberg, 1974) removes all organic substances and gives an ash-free weight. To circumvent the problems of calculating oligochaete production from field surveys, where age-determination is not possible and species identification too tedious or not feasible (immature worms), mortality and productivity may be described as losses between successive sample periods and annual differences in standing stock (Standen, 1973). Erman & Erman (1975) sorted their populations of (terrestial and freshwater) oligochaetes in length classes and determined production by summing “losses” between successive classes multiplied by the number of length classes. Another method of assessing net production is measurement of respiration. The various techniques and their application for oligochaetes (Cartesian diver method, microflow electrodes) are summarized and critically reviewed on the basis of his own studies by Lassèrre (1976). CULTURE METHODS FOR MARINE OLIGOCHAETES Use as fish food necessitated the early development of culture methods for limnetic oligochaetes (La Rue, 1937; Lehmann, 1941; Timm, 1972), whereas some culture experience on their marine relatives has only been made with E. albidus which also occurs in soils and freshwater habitats (Blount, 1937; Le Ray & Ford, 1937; Loosanoff, 1937; Timm, 1972) and is widely cultured on an oat or lettuce diet for aquarium fish. Possible subspecific physiological differences within Enchytraeus populations, however, made the authors feed animals from the shore with decaying wrack (Fucus) buried in sea sand from the same habitat. Fresh thalli have to decompose for a while before they will be accepted by the worms (Schöne, 1971). In principally the same culture conditions Lumbricillus lineatus can be bred through several generations in closed Petri dishes to prevent evaporation and increase of salinity (Giere & Hauschildt, 1979). Other larger sublittoral species (e.g. L. rivalis) can be kept in the same way (Reynoldson, 1947b). Bonomi (pers. comm.) keeps limnetic oligochaetes on moist, sterile sand with chopped lettuce or spinach from deep-frozen packages as food, a method which has been

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developed by Åkesson (1975) for polychaetes and which is certainly extendable to marine oligochaetes. Aeolosoma hemprichi from slightly brackish environments was successfully cultivated in jars on a diet of the alga Dunaliella salina (Eick, Bresk & Spittler, 1973). For breeding terrestrial enchytraeids, Christensen (1956) designed ‘glassslide chambers’, batteries of which he placed in Petri dishes. Besides the advantage of saving space he could, thus, conveniently control the worms and their cocoons between the slides under the dissecting microscope. All attempts to apply this elegant set-up to eulittoral marine species have, however, failed because after a while, the animals regularly suffered from oxygen deficiency and development of hydrogen sulphide, probably due to the high moisture content. Aquatic oligochaetes can be well maintained in ‘water tables’ containing jars with natural sediment in which water is continuously renewed by a dripping water supply (see Johnson & Brinkhurst, 1971). For successful breeding it seems necessary to limit the number of individuals per culture dish to not more than about 100 (Timm, 1972; Hauschildt, pers. comm.). If excessive bacterial and fungal growth impairs the culture medium, treatment with antibiotics (e.g. tetracycline) is advisable (Springett, 1964; Latter, 1977; Giere & Hauschildt, 1979), but their combination and amount has to be adapted to the specific conditions. Cultivation of marine oligochaetes which mainly feed on bacteria, microfungi, and microalgae (diatoms) (see p. 207), is considerably more difficult. Although it is feasible to keep fairly dense populations in sand from their original habitat for months, regular and controlled breeding through longer periods has not yet been achieved, probably due to lack of adequate food. Here, further studies especially with cultures of marine bacteria on seawater agar (Albert, 1975; Tietjen & Lee, 1975, 1977) are highly desirable.

FAUNISTIC AND ZOOGEOGRAPHICAL ASPECTS Oligochaetes, primarily a limnetic group, have spread both into terrestrial and marine habitats. Within the Oligochaeta, the Lumbricomorpha (the former Opisthopora) and the Prosopora are almost totally restricted to terrestrial and freshwater habitats. Only in the family Megascolecidae have some littoral marine species developed: Microscolex and Plutellus sp., occurring especially on temperate and polar islands of the southern hemisphere, and Pontodrilus, known from several tropical and subtropical shores of the Indian and Atlantic Ocean. Hence, almost all marine and brackish-water genera and species belong to the order Naidomorpha (the former Plesiopora), predominantly to the families Tubificidae and Enchytraeidae. The majority of marine oligochaetes is represented by the Tubificidae (Table I). At present, 17 genera are exclusively known from the marine environment (in enchytraeids only one). The bulk of these genera belongs to the subfamily Phallodrilinae (at present 12 genera: Phallodrilus, Aktedrilus, Spiridion, Adelodrilus, Jamiesoniella, Bathydrilus, Peosidrilus, Bacescuella, Inanidrilus, Bermudrilus, Uniporodrilus, and Coralliodrilus), recently revised by Erséus (1980a). Most of them display a characteristically “hollow” prostomium containing relatively few scattered coelomocytes. The increase in marine species has mainly been to this subfamily. The only known

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TABLE I Occurrence of some important naidomorph oligochaete families in the marine environment

Family Aeolosomatidae

Total number of genera

Genera with <50% marine spp.

Exclusively marine genera

2–4

–

–

26–30

2*

–

Tubificidae

45

2·

17

Enchytraeidae

24

2

1

Naididae

* Majority of species abundant in brackish water.

phallodriline freshwater species, Phallodrilus hallae (Cook & Hiltunen, 1975) can be considered as a re-immigrant to the limnetic environment, but its systematic position is to be discussed (Brinkhurst, pers. comm.). Within the Tubificinae, the newly erected genera Tubificoides (Holmquist, 1978) and Heterodrilus (best known species: H.arenicolus) are exclusively marine. Most of the Tubificoides-group formerly belonged to the marine representatives of the genus Peloscolex, recently revised by Holmquist (1978, 1979). The marine species P.benedeni is now placed in the genus Edukemius (E.benedii). As this species is the oligochaete best known to many marine ecologists, we shall (in this review) use the synonym Peloscolex benedeni. Some species of the genus Tubifex are also abundant in marine and brackish-water environments (e.g. Tubifex costatus and T.pseudogaster). The genera Monopylephorus (Rhyacodrilinae) and Limnodriloides (Aulodrilinae) are predominantly marine. The aulodriline genus Thallassodrilus is exclusively found in the marine milieu, as is the genus Clitellio (Clitellinae). Most marine Enchytraeidae belong to the genera Lumbricillus and Marionina which populate predominantly marine and brackish intertidal and supralittoral habitats and only rarely freshwater and terrestrial biotopes. Also the common genus Enchytraeus includes marine, limnetic and terrestrial forms. Even some members of the limnetic genera Cernosvitoviella, Fridericia and Achaeta populate marine or brackish beaches. On the other hand, the sublittoral genus Grania is exclusively marine. The naidid genera Paranais and Amphichaeta are mainly distributed in marine and brackish habitats. The cosmopolitan freshwater species Nais elinguis occurs frequently in brackish water. The monotypic genus Wapsa (Marcus, 1965) is reported from Brazilian brackish environments. The Phreodrilidae which are restricted to the southern hemisphere, have some marine littoral representatives (e.g. Hesperodrilus litoralis Michaelsen, 1924). From the Aeolosomatidae, only one truly marine species, the mesopsammic Aeolosoma maritimum (Westheide & Bunke, 1970), is at present known. Several freshwater species, e.g. A.hemprichi and A.litorale are frequent in brackish water.

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Considering the rapidly increasing number of newly described marine species, it would be premature to discuss their zoogeography in detail. Many species are only known from their type locality (especially in the recently described phallodrilines; Erséus, 1980a), which only indicates that Europe and North America are at present the centres of oligochaete research. Oligochaetes from the Pacific American coast are still little known and information from other continents is extremely fragmentary. Species from South America are only mentioned in a few papers (e.g. Marcus, 1965; Righi & Kanner, 1979). Scant reports on Asian marine oligochaetes are mainly given by Shurova (1974, 1977a,b) from the Soviet Pacific coast. McLachlan and co-workers found numerous interstitial oligochaetes on South African beaches (e.g. McLachlan, 1977b). Some of their specimens, kindly lent to the senior author, could be tentatively identified as new tubificid species. Recently, numerous new tubificids and enchytraeids were described from the southern tropical hemisphere (e.g. Jamieson, 1977; Erséus, 1981—Great Barrier Reef). The fact that oligochaetes have now also been discovered in abyssal and hadal depths underlines the need for further investigations.

TABLE II Geographical distribution of some common marine oligochaetes: Atl., Atlantic; Pac., Pacific; Ind., India; Cosmopol., cosmopolitan; LT, littoral; ET, eulittoral; SP, supralittoral ; ST, sublittoral; M, marine; B, brackish water; F, fresh water; T, terrestrial; +, present

N. S. America America Species

Africa

Asia

Habitat Europe Atl. Pac. Atl. Pac. Atl. Ind. Ind. Pac. Cosmopol.

Naididae Amphichaeta sannio

LT-ST, B

+

Paranais litoralis

LT-ST, M, B

+

P.frici

LT-ST, B, F

+

Nais elinguis

LT-ST, B, F

+ +

+

+

+ +

Tubificidae Tubifex costatus

ST, M, B

+

T.pseudogaster

ST, M, B

+

+

Peloscolex

ST, M,

+

+

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B

Tubificoides heterochaetus

ST, M, B

T.gabriellae

ST, M, B

+

+

T. nerthoides

ST, M

+

+

T.apectinatus

ST, M

+

+

Limnodriloides monothecus

ST, M

+

+

Clitellio arenarius

ST, M, B

+

+

Aktedrilus SP-ET, monospermathecus M, B

+

+

+

215

+ +

Monopylephorus rubroniveus

ST, B

+

M.irroratus

ST, B

+

M.parvus

ST, B

+

Enchylraeus albidus

T-SP, M, B, F

+

Lumbricillus lineatus

SP-ET, M, B

+

+

L.pagenstecheri

SP-ET, M, B

+

+?

Marionina subterranea

SP-ET, M, B

+

+

M.spicula

SP-ET, M, B

+

+

M.southerni

SP-ET, B

+

+

Enchytraeidae

+

+?

+

Hence, such statements as Stephenson’s (1930) that Enchytraeidae are “certainly relatively uncommon” and “scarcely found endemic in the tropics and southern hemisphere”, or that tubificids are “less common in the hotter regions of the globe” today must be considered unjustified. It is likewise quite speculative that Oligochaeta should have originated from cooler climates (Timm, 1980). Besides scarcity of investigations, also technical and identification problems account for the little known status of marine oligochaetes: most species inhabiting sublittoral bottom are <10 mm in length and easily overlooked or discarded when sieving for

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macrofauna. Even the frequent littoral and intertidal species are often merely reported as “Oligochaeta” due to severe identification problems. Often only the occurrence of Peloscolex benedeni, easily identifiable by its thickly papillated body wall, has been noticed by people who are not experts on the Oligochaeta. Marine tubificids are also often mistaken for capitellid polychaetes (especially Capitella capitata) when associated in the same habitat. As a detailed general analysis of the zoogeography of Oligochaeta has been recently published by Timm (1980), and for the reasons discussed above, the zoogeographical data compiled here (Table II) will only consider reliable reports on some common species found on both sides of the Atlantic Ocean. Among the Naididae, Paranais litoralis occurs on both sides of the North Atlantic, and has also been found in Africa and Asia. P.frici, so far only known from European rivers (e.g. Rivers Elbe and Donau) and brackish waters (Baltic Sea), was recently detected in the Fraser Estuary (British Columbia) (Chapman & Brinkhurst, 1980). The brackishwater species Amphichaeta sannio, known from a variety of European localities (e.g. Baltic Sea), is supposed to occur in China (see Brinkhurst & Jamieson, 1971). The tubificid Tubifex costatus is only known from European coastal waters, while Aktedrilus monospermathecus (?), Tubificoides heterochaetus, Clitellio arenarius, and Peloscolex benedeni are amphi-North Atlantic. Tubifex pseudogaster is frequent on both sides of the North Atlantic and has also been found on the Pacific coast of British Columbia, although this may have been an assemblage of sibling species (Brinkhurst & Baker, 1979). The occurrence of sibling species is also under discussion for Tubificoides gabriellae which is found on the Atlantic coast of North and South America as well as on the Pacific coast of the United States. Monopylephorus rubroniveus, M. parvus, and M. irroratus are regarded as brackish-water cosmopolitans (Brinkhurst & Jamieson, 1971; Brinkhurst & Baker, 1979). Limnodriloides monothecus, Tubificoides apectinatus, and T.nerthoides occur on the east and west coast of North America. The littoral and terrestrial enchytraeid Enchytraeus albidus is cosmopolitan. Lumbricillus lineatus, abundant on European and North American shores, is perhaps also circum-mundane. L.pagenstecheri, found in Europe including Iceland, is also supposed to occur in Greenland, Canada, and Alaska (see Erséus, 1967b). The meiobenthic species Marionina southerni is amphi-North Atlantic. M.subterranea has been found on both North Atlantic coasts as well as along the North American coast of the Pacific (Lassèrre, 1971a; Cook & Brinkhurst, 1975; Locy, pers. comm.; Giere, in prep.).

REPRODUCTION, LIFE CYCLE AND POPULATION STRUCTURE Studies of the life history of oligochaetes are aggravated by a number of complicating factors which explain in part the scarcity of data on marine species. (1) Absence of discrete age classes. (2) Neither body size nor number of segments are precise features for determination of the stage of maturity; they can only be grouped statistically by division into lengthfrequency or width-frequency classes (e.g. Ladle, 1971b; Hunter & Arthur, 1978). (3) Examination of the stage of maturity involves time-consuming microscopical

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examination of each specimen. (4) In a mixed oligochaete population, the cocoons of the different species are not identifiable by morphological features. Only hatching in the laboratory can solve the question of species. (5) Culturing aquatic oligochaetes, especially tubificids and naidids, often fails because of the long time needed for life-history studies; this failure is partly due to nutritive problems and to chemical changes in the culture water. Naidids particularly are highly sensitive to alterations in water composition; results have been obtained in enchytraeid cultures (see p. 191). AEOLOSOMATIDAE In the Aeolosomatidae asexual reproduction occurs principally by fission. The animals produce stolons similar to those of syllid polychaetes. The position of the fission zone is flexible (Bunke, 1967). In Aeolosoma maritimum the stolons do not separate anterior to the Xth or XIth segment of the first zooid (Westheide & Bunke, 1970). Although genitalia can develop in Aeolosomatidae, sexual reproduction rarely occurs, as the eggs atrophicate at a certain stage and all sexual differentiations are re-absorbed. Only in the freshwater species A. quaternarium have egg deposition and development of embryos been observed. NAIDIDAE In naidids, paratomy by budding off zooids represents the normal type of asexual reproduction while architomy is rather rare; only few species (e.g. Allonais spp.) merely fragment. The fission zone seems in some species to be genetically fixed, e.g. segment XXXVIII in Dero spp. and XXXV in Ophidonais (Dehorne, 1916). In other genera, the position of the fission zone was influenced by extrinsic factors like temperature and nutrition, as described for Pristina longiseta by Cleave (1937). The process of budding in Nais elinguis represents the typical type for naidids according to Herlant-Meewis (1946): a chain of zooids is normally formed by three stolons, one anterior, one intermediate, and one posterior. The separation of the posterior zooid takes place when 15–16 segments are developed. In some cases, a new zooid can be formed on the posterior stolon before budding off. After the separation of the posterior zooid, a new stolon appears between the anterior and the former intermediate zooid, which now represents the new posterior zooid. The fission-rate generally increases with rising temperature up to an individual optimum temperature (Cleave, 1937; Poddubnaya, 1968). Optimum temperature for Chaetogaster diaphanus is ≈20°C (Poddubnaya, 1968), for Amphichaeta sannio, Koene (1981) tentatively found 12–18°C. Loden (in prep.) could stimulate the budding activity of cultured Pristina spp. by an artificial rise of the sodium chloride content of the water. McElhone (1978) stated the number of buds to be significantly correlated with food availability. This was also demonstrated by Poddubnaya (1968) for Chaetogaster diaphanus; with insufficient or poor quality of food, this species reduced growth rate of the zooids. Here, budding occurred, on the average, every second day. McElhone (1978) reported a doubling time in a laboratory population of Nais

Interlinking of physical

692

pseudoobtusa within 4.2 days but it took 26 days for the field population. Amphichaeta sannio budded off zooids at the most every second day in field populations from the Dollard Estuary, in the laboratory, however, this span was extended to ≈10 days in similar temperature conditions (Koene, 1981). By intensive budding activity, naidid populations can multiply within a short time which always leads to an enormous increase in individual numbers in natural populations. In the Schlei-Förde (western Baltic), Paranais litoralis started budding in March–April up to a maximum in June. The number of individuals increased from ≈1000·m−2 (beginning of March) to ≈18 000·m−2 (June). In summer, its abundance is evidently diminished to 4000–6000·m−2 (Fig. 2). Amphichaeta sannio from the Schlei-Förde (Fig. 2) and A.leydigii from the tidal freshwater flats of the Elbe Estuary (Pfannkuche, 1980c) are characterized by two annual periods of budding activity (early spring and autumn) which coincide with blooms of benthic diatoms. Nearly all common naidid species in the marine environment develop population maxima by means of asexual reproduction at times of moderate temperature and good food supply, i.e. mainly in spring (Heip, 1971; Watling, 1975; Learner, Lochhead & Hughes, 1978; Pfannkuche, 1979; Koene, 1981) or in autumn (Kendall, 1979). Many authors, e.g. Pfannkuche (1979) and Koene (1981) stress this dominance of asexual reproduction for marine and brackish-water species, and McElhone (1978) for freshwater species. There is little evidence that sexual reproduction has an important function in the growth of naidid populations. Sperber (1948) reported only very few findings of sexually mature individuals. Hughes (1975, in Learner et al., 1978) found only three species out of seven to reproduce sexually during one season in a Welsh mountain stream. There is, apparently, only one sexual generation a year with the adults dying soon after breeding (Learner et al., 1978). Most authors report the occurrence of mature specimens during times of active asexual reproduction and population peaks. In the Paranais litoralis population of the Schlei-Förde, up to 20% mature and reproductive specimens were found during the spring maximum (Fig. 2), although mature individuals also occurred in October (Pfannkuche, 1974). Similarly, mature and reproductive Nais elinguis were only abundant during the spring maximum, both in the Schlei-Förde (Pfannkuche, 1979) as well as in freshwater tidal flats of the Elbe Estuary (Pfannkuche, 1981). In Amphichaeta from the Schlei-Förde, mature individuals (<5% of the population) were only encountered during the spring maximum, whereas in the Dollard

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Fig. 2.—Seasonal abundance of naidids in various breeding classes in the αmesohalinicum of the Schlei-Förde (western Baltic), 1974 (redrawn from Pfannkuche, 1979).

Interlinking of physical

694

Estuary, the same species was found to reproduce sexually in autumn (Koene, 1981). From ponds, small streams and ephemerous waters, larger numbers of mature specimens were reported. Loden (in prep.) found 52.8% Nais variabilis and 62.1% N. communis in various stages of sexual maturity in some Indiana (U.S.A.) streams. A large number of mature Paranais litoralis (>50%) was also found in some temporary brackish-water pools at the western Baltic (Pfannkuche, unpubl. data). The life-cycles of naidids exhibit a high plasticity depending on the local habitat conditions. Several factors, such as temperature, changes in water chemistry, and nutrients, influence the rate of asexual reproduction. The factors stimulating sexual maturity are not sufficiently known. There is some evidence (Stolte, 1921, 1969) that maturation is also controlled by certain changes in water chemistry and food composition. Maturation may also be stimulated by intrinsic factors triggered by population peaks (see above). Interrelations between population density and breeding activity were also found in Tubificidae (Poddubnaya, 1980) and in Enchytraeidae (Christensen, 1973). Species inhabiting relatively stable habitats become sexually mature in relatively small quantities, as was demonstrated both for marine (Pfannkuche, 1979) and freshwater (Learner et al., 1978) species. In astatic locations with unpredictable, heavy changes of the ambient milieu and food supply, the majority of the naidid population seems to become sexually mature as cocoon deposition ensures better survival during periods of extreme environmental stress (Loden, in prep.). TUBIFICIDAE Asexual reproduction, although described for some freshwater representatives (Hrabé, 1937), is negligible in Tubificidae. Life-history studies have, so far, only been conducted on central and northern European shores (mainly North Sea and Baltic) in Tubifex costatus (Brinkhurst, 1964; Pfannkuche, 1979; Birtwell & Arthur, 1980) and in the meiobenthic species Aktedrilus monospermatheeus and Spiridion insigne (Pfannkuche, 1980b). For analysis of their life cycles, tubificids are generally classified as follows: (1) immature; (2) mature (worms with fully developed genitalia); (3) breeding or reproductive (worms containing spermatophores and eggs); (4) post-breeding (adult specimens lacking spermatophores and eggs, having re-absorbed their genitalia, but retained penis sheaths). A post-breeding stage was described in Tubifex costatus by Birtwell & Arthur (1980) and in Peloscolex benedeni by Hunter & Arthur (1978). The presence of a morphologically distinguishable post-breeding stage was also found by Poddubnaya (1980) in Limnodrilus hoffmeisteri and Tubifex tubifex. So far, all investigations on the marine tubificids mentioned above, reveal a similar picture of breeding behaviour. The worms reach sexual maturity in late winter or early spring and reproduce in spring and early summer. A typical example is given in Figure 3 for T.costatus and T.pseudogaster from the α-mesohaline part of the Schlei-Förde (after Pfannkuche, 1979). In autumn, the population of T.costatus consisted nearly entirely of immature individuals. In late winter, mature worms became abundant. Their proportion increased from 20% in January to ≈40% in February, but specimens with spermatophores

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were still rare (<5%). Reproductive worms first appeared in quantity in March (≈20%). In April, the number of breeding worms was ≈35%, whereas the mature individuals were <20%. Altogether, from March to May about half the population was in a breeding or mature stage, while the other half remained immature. Cocoons were found from the middle of March to the end of July, but they could not be distinguished from those of T.pseudogaster and Peloscolex benedeni which occurred in the same

Fig. 3.—Occurrence of breeding classes in Tubifex costatus and T.

Interlinking of physical

696

pseudogaster in the α-mesohalinicum of the Schlei-Förde (western Baltic), 1974 (redrawn from Pfannkuche, 1979).

habitat. From June onwards, after the hatching of a new generation, immature specimens clearly dominated with 90% in July and remaining between 90 and 100% until the end of the year. Brinkhurst (1964) observed an analogous cycle in a monospecific population of Tubifex costatus in Lancashire (U.K.), but maturation and cocoon deposition was somewhat delayed compared with the situation in the Schlei-Förde. Birtwell & Arthur (1980) found the maximum of mature T.costatus in the Thames Estuary (U.K.) in December and January; breeding worms predominated from the end of February to early May. Mature T.pseudogaster in the Schlei-Förde could be found throughout the year, but <5% occurred between August and December. In late winter, their number increased rapidly (Fig. 3). At the end of February, 35% were in mature stage. Breeding worms were abundant from March onwards but, unlike T.costatus, the number of mature worms was not exceeded by the reproductive ones until May. In T.pseudogaster, one part of the population also remained immature during the breeding season. Cocoons were supposed to be produced between May and July. Hence, except for slight differences in the maximum of breeding activities (May-June), the reproductive cycle of T. pseudogaster is similar to that of T.costatus. Mature Peloscolex benedeni in a nearly monospecific population from the Thames Estuary were reported by Hunter & Arthur (1978) to occur in December, with a maximum in late February. In the Schlei-Förde, maturation and reproduction was observed to start one month earlier (Pfannkuche, 1979). In the Thames population, breeding worms increased in number from April onwards and reached a maximum in early June. Maximal cocoon deposition was observed in July. The recruitment rate of newly hatched worms increased in June and reached a maximum in July. Breeding activities ended in late August. In Norwegian shallow waters, a corresponding reproductive period in the warmer months was reported by Erséus (1976a). Worms of post-breeding stage followed the breeding stage, but their proportion declined with maturation of a new generation. The data of Hunter & Arthur (1978) do not answer the question whether the post-breeding worms die after a certain time or take part in a second breeding season. Growth rate in the Thames population seems to be slow, as immature worms always formed more than half of the population even during the breeding period. The interstitial supralittoral species Aktedrilus monospermathecus and Spiridion insigne have seasonal reproduction activities similar to those of the above subtidal macrobenthic species (Fig. 4). Maturation in the Schlei-Förde commenced in late winter and reached a maximum in March and April (≈30% mature specimens). In late spring, the number decreased constantly until August. Breeding specimens occurred first in late February; in April and May, their number surpassed the mature ones. Maximal breeding activity was in late spring and early summer and came to an end during August. Even at peak times of reproductive activity, about one half of the population remained in an immature stage. In July, immature worms predominated (≈70%). In autumn and early winter, more than 95% of the population were immature.

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The above life-history studies show, despite some temporal differences, an overall similarity which may be interpreted as an adaptation to shallow-water fluctuations. Subtidal meiobenthic tubificids, living in a more stable environment, seem to have reproductive periods scattered over the whole year (Erséus, 1976a). The differences in occurrence and duration of breeding classes reflect divergences in ambient factors. Growth, reproduction, fertility rates, and maturation time in tubificids are significantly influenced by temperature (Kennedy, 1966; Aston, 1968, 1973a; Thorhauge, 1975, 1976; Wiśniewski, 1976; Poddubnaya, 1980). Thus, development and reproduction are also affected by oxygen content, sediment structure (Palmer, 1968; Aston, 1973a; Hunter & Arthur, 1978), food availability (Kosiorek, 1974; Pfannkuche, 1981; Poddubnaya, 1980), and population density (Poddubnaya, 1980). Despite these modifying influences, the data so far available, seem to indicate a two-year life cycle as the basic type for marine

Interlinking of physical

698

Fig. 4.—Seasonal occurrence of breeding classes in some common interstitial boreal oligochaetes (modified after Lassèrre, 1971b; Pfannkuche, 1980b).

tubificids (Brinkhurst & Jamieson, 1971), which means that the worms, as a rule, do not reproduce before their second year (Brinkhurst, 1964; Hunter & Arthur, 1978; Pfannkuche, 1979, 1980b). This pattern corresponds to conditions found in Lumbriculidae (Cook, 1969a).

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ENCHYTRAEIDAE Although the bulk of enchytraeids reproduce sexually by mutual fertilization, other ways of reproduction are also of relevance. Asexual reproduction by fragmentation is common in several limnetic and terrestrial species, e.g. in Cognettia sphagnetorum (Christensen, 1959; Springett, 1970; Standen, 1973) and Enchytraeus bigeminus (Christensen, 1973). Polyploidy in enchytraeids is frequently encountered (Christensen, 1961) and further investigations may reveal more examples of parthenogenesis in this family. Among the marine species, polyploid caryotypes were found in natural populations of Lumbricillus lineatus, where beside the common diploid type (2n=26), a certain number of triploids may be encountered. The triploid cytotype is distinguishable from the diploid by the absence of seminal vesicles and sperm production (Christensen & O’Connor, 1958; Christensen, 1960, 1961; Christensen, Jelnes & Berg, 1978). Although the triploids reproduce parthenogenetically, their eggs require ‘stimulation’ by sperm from sexually normal diploid worms for successful development (pseudo-fertilization). Even tetra- and pentaploid caryotypes were recently detected in a natural population of L.lineatus (Christensen et al., 1978). The life cycles of field populations in marine enchytraeids are not well known; even from terrestrial habitats only a few studies have been done (e.g. Springett, 1970; DózsaFarkas, 1973). As culturing is much more successful in enchytraeids than in tubificids and naidids, life-history studies, based mainly on laboratory data, are available for some ubiquitous marine species. Since the classical work of Reynoldson (1939b, 1943) on “L.lineatus” (L. rivalis), several investigations have been made on the life histories of L.rivalis (Williams, Solbé & Edwards, 1969; Kirk, 1971; Learner, 1972), L.lineatus (Hauschildt, 1978; Giere & Hauschildt, 1979) and Enchytraeus albidus (Reynoldson, 1943; Ivleva, 1953; Schöne, 1971). A transformation of data on Lumbricillus rivalis from sewage percolating filters to the situation on the seashore seems particularly problematical since Christensen & Jelnes (1976) described the existence of at least two ecologically separated sibling species in this morphologically identical assemblage. The first group populates the seashore, the second is abundant in fresh and oligohaline waters (see p. 273). The growth rate, fertility, and reproduction biology of the enchytraeids from marine wrack beds and sewage percolating filters is characterized by an extreme flexibility as an adaptation to changes in temperature, salinity, and nutrition (Table III). The divergent recordings in L.rivalis and Enchytraeus albidus in Table III may refer to differing temperature optima, which depend on local habitat conditions. This was demonstrated by Hauschildt (1978) for Lumbricillus lineatus. Depending rather on temperature than on salinity, its embryonic development lasts from ≈6 days at 26°C, to 11 days at 13°C. Even at the same temperature (13°C), it can vary from 9.6 to 13.1 days. Similarly, the mean generation time is 99±44 days (13°C). As a reaction to extreme stress conditions (high salinity), embryogenesis can also be considerably delayed due to retardation periods during certain developmental phases. In times of starvation, L.lineatus can even stop maturation or egg production and degenerate genitalia. On the other hand, under good conditions it can develop within a short time. Reaching maturity within 30–40 days after hatching, it starts an intensive cocoon production (on average 1.1 cocoon with 10 eggs

Interlinking of physical

700

per cocoon every week) for 12–18 weeks. Hence, about 190 hatched worms could be reared per worm and reproductive period. In cultures, 4–6 generations (from cocoon to cocoon) per yr are not unusual (Giere & Hauschildt, 1979). The ability to regenerate sexual organs after reproduction and to substantially take part in a second breeding period, as to be judged from the cultures (Hauschildt, pers. comm.), underlines the extensive reproductive potential of this euryvalent species. Participation in a second breeding season after a total regeneration of genitalia and retrojuvenation (“ananeosis”) was also found by Dózsa-Farkas (1973) in the terrestrial species Stercutus niveus.

TABLE III Reproduction of three enchytraeids in relation to temperature: (modified from Hauschildt, 1978)

Species

Lumbricillus lineatus

Lumbricillus rivalis

Enchytraeus

Reference Hauschildt, Temp. Reynoldson, Learner, Kirk, Temp. Reynoldson, Learner 1978 (°C) 1939b, 1943 1972 1971 (°C) 1939b, 1943 1972 No. of eggs per

9.8 (2–26)

13

7

17 0–39

10

cocoon

7.9 (3–19)

26

7

13

20

4–5

Egg fertility (%)

80*

13

74

52*

71– 91

10

83.2

65*

26

58

41*

31– 92

20

9–12

13

10

9

17.4

10

14

2

4–7

26

6

8

20

10.3

1

75 (30.5– 135)

13

62.3

33

15

26.6

4

51 (24–115)

26

50.5

26

20

28.4

2

Generation period

85.5 (39– 147)

13

72.2

45

15

40.5

6

from cocoon to cocoon (days)

56 (28.5– 120)

26

56.1

34

20

38.7

6

Embryonic develop ment (days) Period of breeding (days)

* Hatching fertility.

63

47

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227

The flexible adaptability of development and production to ambient conditions, emerging from laboratory data, renders it difficult simply to use laboratory figures for calculations on population dynamics in the field. The wide variation in maturation time and cocoon production, already apparent in long-term cultures of Lumbricillus lineatus (see above) originating from one population, might be even wider if field and culture data from different populations were compared. In laboratory cultures of Enchytraeus albidus, egg number in cocoons and number of hatching juveniles was found to be considerably higher than in the field (Reynoldson, 1947a,b). A high adaptability to ambient conditions was also monitored by Wiśniewski (1979) for the fecundity of limnetic tubificids. Although the daily egg production decreased with depth of cocoon deposition in the mud, overall productivity could not be related to depth, since breeding periods were found to be extended in populations from greater depths. On the other hand, a large number of eggs and newly hatched juveniles was eliminated by a mortality rate of ≈70%, the depletion being most severe in the surface layers and least in 2–5 cm depth (data with predators excluded). In lumbricid cultures, the number of cocoons and of eggs in them increased about three times, whereas maturation time was cut down to even 10–20% of the field values (Stolte, 1969). The life history of meiobenthic enchytraeids is almost totally unknown. Although many species (e.g. Marionina spp.) produce only 1–2 eggs at a time, repeated cocoon deposition by one individual (according to Lassèrre, 1975a: 2–4 cocoons) and short incubation periods (10–15 days) render rich stocks of meiobenthic enchytraeids in the marine eu- and supralittoral (Giere, 1975) which leads to the assumption of short generation cycles. Lassèrre (1971c, 1975a) found mature Marionina spp. in the Bay of Arcachon (France) during spring and autumn (see Fig. 4); he assumed the existence of 2– 3 generations a year in most populations of Marionina. MEGASCOLECIDAE Sexually mature Pontodrilus bermudensis in quantity appeared in Indian waters at the end of the southwest monsoon (Subba Rao & Ganapati, 1975). Cocoon deposition started in November and continued until May, with a peak in February and March. Also Stephenson (1915) reported the occurrence of mature P.bermudensis in late winter and early spring. POPULATION STRUCTURE Population dynamics of oligochaetes differ largely depending on their family affiliation and habitat conditions. In naidids, the population structure is generally characterized by heavy seasonal fluctuations in abundance which may lead to an almost total breakdown of the population or even to a temporal disappearance. A typical example of seasonal variations of the ubiquitous marine and brackish-water species Paranais litoralis, Amphichaeta sannio, and Nais elinguis is illustrated in Figure 5. At Karschau (αmesohaline part of the Schlei-Förde, western Baltic) abundance is <1000 ind.·m−2 from end of November until beginning of February. Towards spring, numbers increase, due to a rise in the Amphichaeta sannio

Interlinking of physical

702

Fig. 5.—Seasonal abundance of naidids in the α-mesohalinicum of the SchleiFörde (western Baltic) (after Pfannkuche, 1980c).

population which attains its maximum in April (21 000·m−2) and then drops again to ≈1000·m−2 in July and August. The increase in this species coincides with the development of its diet, benthic diatoms (see p. 207). The number of Paranais litoralis and Nais elinguis increase continuously from the middle of March reaching a maximum in June (17 500 Paranais litoralis·m−2 and 4500 Nais elinguis·m−2) resulting in a total of naidids of 34 500·m−2 in June. Shortly after this peak, the naidid population drops to only 10 000·m−2 in July. This steep decline is seen for both Paranais and Nais and continues until November. Amphichaeta sannio, in turn, develops a second maximum of about 8000·m−2 in September. Two annual maxima, in early spring and in autumn, were also observed in a freshwater tidal flat population of A.leydigii in the upper Elbe Estuary by Pfannkuche (1981). Erratic oscillations in the development of Paranais litoralis with two main peaks (spring and late summer) were found in the polyhaline (20–30‰ S) intertidal flats of the Forth Estuary (Bagheri & McLusky, in press). In the same area, Amphichaeta sannio showed only one annual maximum, varying from April to August. That these population outbursts are mainly due to intensive budding is deduced from the biomass curves (Bagheri & McLusky, in press) which often dropped at times of maximal abundance. Heavy fluctuations in the abundance of Paranais litoralis with peaks in spring were also described by Watling (1975) and Kendall (1979). Nais elinguis and Paranais frici

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exhibited a similar course of seasonal fluctuations in the freshwater tidal flats of the Elbe Estuary (Pfannkuche, 1981). Seasonal variations in abundance were also reported for several freshwater species by other authors (e.g. Learner et al., 1978). The period of maximum abundance in a species differs locally depending on the habitat conditions. Nais elinguis, in a Welsh stream, was found most numerous in winter and spring (Ladle, 1971b), but in the phreatic water of river bank gravels it was most frequent in summer (Ladle, 1971a). According to Learner et al. (1978), N.elinguis is a spring species, which is confirmed by the above data from the Schlei-Förde, but it is clearly variable in its seasonal occurrence. An aspect of particular interest is the complete disappearance of a species mainly in winter. This was described for Paranais litoralis by Watling (1975), can be deduced from the results of Bagheri & McLusky (in press) for this species, and also for Amphichaeta sannio and has been reported for several freshwater species by Ladle (1971a) and Mason (1977). Learner et al. (1978) suggested as a possible explanation that the worms penetrate more deeply into the sediment thus avoiding adverse surface conditions. But studies on the vertical distribution of naidids (Hughes, 1975, in Learner et al., 1978; Pfannkuche, 1980c) gave no evidence to support this assumption. According to another possible argument, the worms over-winter in any form of resistant stage (Learner et al., 1978), e.g. in cocoons. In the Schlei-Förde, a complete disappearance of naidids was not observed. It seems more likely that in many cases naidid abundance becomes so low that they are only occasionally caught. The survival of only a few individuals during times of adverse environmental conditions would be sufficient for a new rapid increase of the population within a short time, as was demonstrated above when discussing the enormous reproductive capacity of naidids by means of paratomy. Little is known about causative factors for the seasonal changes in naidid populations. As with the reproduction pattern (see p. 187). they seem to be stimulated by changes in water chemistry, temperature, oxygen regime, and food availability. The population dynamics in (larger) tubificids differ completely from the naidids. Tubificid populations maintain a big standing stock throughout the year, although considerable seasonal variations in abundance were also described by some authors (e.g. Poddubnaya, 1959; Hunter & Arthur, 1978; Birtwell & Arthur, 1980; Bagheri & McLusky, in press). This relative constancy is mainly due to the reproductive pattern of tubificids. Many species like Tubifex costatus and Peloscolex benedeni have a two-year cycle and only a part of the population is in a reproductive stage during one breeding season while the other part remains immature and breeds during the next season. The time span of breeding activities extends, although dependent on the ambient conditions, over some months, so that the newly hatched generation is gradually recruited over a longer period. The data on the population dynamics of tubificids differ markedly when comparing those from permanently submersed bottoms of the Baltic Sea with those from the intertidal of estuaries. Living conspecifically in the α-mesohalinicum of the Schlei-Förde, (Pfannkuche, 1979, 1980c), the standing stock of Tubifex costatus and T.pseudogaster (Fig. 6, Curves I and II) varied not more than ±10% of the annual mean. No distinct seasonal maxima of abundance were noticed. A corresponding range of variation was found in a monospecific T.costatus population in the same region. From the intertidal

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704

zone, Brinkhurst (1964), studying the Irish Sea shores at Hale, U.K., reported, however, a seasonal variation of ±25% of the annual mean in a monospecific T.costatus population (Fig. 6, Curve III). The increase in April cannot be referred to an enhanced recruitment of young worms, as cocoons were not abundant before May. It is more likely that these fluctuations can be referred to non-homogenous distribution patterns which could not be covered by duplicate samples. Hunter & Arthur (1978) studying a

Fig. 6.—Seasonal abundance of some marine tubificids: I, Tubifex costatus, αmesohalinicum, Schlei-Förde (Western Baltic), 1974 (Pfannkuche, 1979); II, T.pseudogaster, α-mesohalinicum, Schlei-Förde, 1974 (Pfannkuche, 1979); III, T.costatus, Irish Sea, Hale, U.K., 1961–1962 (Brinkhurst, 1964); IV, Peloscolex benedeni, Thames Estuary, North Sea, 1972–1973 (Hunter & Arthur, 1978); V, T.costatus, Thames Estuary, 1970–1971 (Birtwell & Arthur, 1980).

monospecific intertidal population of Peloscolex benedeni in the Thames Estuary also found heavy fluctuations in abundance. Again, the increase of worms from from January to April could not be explained by recruitment

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through breeding activities, as newly hatched worms first occurred in quantity in May. The peak in 1972 was reached in August (Fig. 6, Curve IV), afterward which the number decreased again to in December. In 1973, the number of P.benedeni again increased rapidly after May due to the recruitment by the newly hatched generation reaching its maximum of 142×103·m−2 again in August. The marked increase in absolute number (nearly 100% in 1973 compared with 1972) was explained by the authors as due to a rise in fertility which in 1973 was 1.4 times higher than in 1972. Considerable discrepancies in population development between stations and years are also characteristic for the stocks of P.benedeni in the Forth Estuary (Bagheri & McLusky, in press). While usually a spring (April) and a summer (July–August) maximum in numerical abundance was found, clear recruitment phases could not be determined. From the intertidal of the Thames Estuary, Birtwell & Arthur (1980) also reported heavy seasonal fluctuations in an almost monospecific population of Tubifex costatus (Fig. 6, Curve V). During their studies, a marked drop from 400×103·m−2 to was found in November. Population density increased again to before the beginning of the breeding season and reached to a maximum of 601×103·m−2 in August, which was referred to the recruitment of the newly hatched generation. From August to September, a drop to nearly 100×103·m−2 was noticed. As a possible explanation, the authors assumed erosion of the upper sediment layers by high river flows or wave action and subsequent transportation of the juvenile tubificids which predominate just within the surface sediments. The population dynamics of tubificids appear closely related to the habitat conditions. Habitats without any abrupt changes in abiotic environmental factors and with a relative continuous food supply mainly from allochthonous processes (e.g. sewage waters) are characterized by high abundances and a relative stable standing stock. Rigorous changes in temperature, oxygen regime, and in the amount of sedimented particles have a strong influence on growth rate, fertility, and mortality of a population with the consequent marked fluctuations. For the meiobenthic Aktedrilus monospermathecus, Hulings (1974) found, on Lebanese beaches, spring peaks in March to May, relatively high densities in autumn, and minima in summer and winter. As a whole, population dynamics in enchytraeids can only be estimated from fragmentary observations. It seems likely that eu- and supralittoral enchytraeid populations fluctuate more strongly in their abundance due to the astatic and unpredictable conditions in their habitats (see p. 251 ff) than due to regular seasonal peaks. Short-term oscillations of abiotic factors are mostly endured by a wide tolerance capacity (Giere, 1977; Giere & Hauschildt, 1979) and are partly avoided by migration (Giere & Pfannkuche, 1978; Meineke & Westheide, 1979; Pfannkuche, 1980b).

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Fig. 7.—Seasonal population fluctuations of Marionina spicula and M. achaeta from the high-water zone of an Atlantic beach near Arcachon (France) in relation to precipitation and salinity changes (modified after Lassèrre, 1971c).

Field data on seasonal fluctuations in marine enchytraeids are, as yet, very scarce. In the supralittoral of French beaches, Lassèrre (1971c) found population peaks of interstitial Marionina spp. in spring and autumn depending on moisture and salinity (Fig. 7). Reproduction almost ceased in cold winters (see Fig. 4). Summarizing our own observations, there is a general depression of population abundance during winter which the worms seem to pass through in submature or juvenile stages. Only Lumbricillus lineatus and Enchytraeus albidus were found to reproduce (at a reduced rate) in the cold season (see also p. 190 ff). This winter depression in enchytraeids is, therefore, not comparable with the breakdown in naidids. In the marine littoral, the remarkable fertility and growth capacity in opportunistic species like Lumbricillus lineatus and the ability to develop several generations per year in many interstitial forms like Marionina spp. (see p. 193) lead to a rapid recovery of population abundance in spring. Seasonal fluctuations in the enchytraeid populations of moorlands were also described by Springett (1970) and Standen (1973).

ABUNDANCE, BIOMASS, AND PRODUCTION ABUNDANCE Oligochaetes are often surprisingly numerous in brackish and marine habitats. Muus (1967) studying some α-mesohaline bights of the Kattegat region, ranked tubificids as the

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most important taxon (except for Nereis diversicolor) amounting to ≈32% of the total macrofauna abundance. The quantitative importance of oligochaetes from various biotopes and geographical regions is demonstrated in Tables IV–VI; although here maximum values are recorded, the numerical dominance of oligochaetes in transects from the supralittoral to the upper sublittoral is obvious. The supralittoral and eulittoral mesopsammal (Table IV) is mainly inhabited by enchytraeids, mostly Marionina spp. The presence of tubificids is usually restricted to some members of the Phallodrilinae (e.g. Aktedrilus monospermathecus, Phallodrilus prostatus). These meiobenthic oligochaete species can attain amazing population densities as is shown by the values of McIntyre & Murison (1973) and Giere (1975) for Marionina subterranea. They are comparable with local maximum aggregations of freshwater tubificids in sewage treated waters as reported by Palmer (1968). Although it seems likely that these maximum values often represent just patchy aggregations, population densities of 50 to 100×103 ind.·m−2 seem to be quite normal for many beaches and sandy shores (cf. Lassèrre, 1967; Pfannkuche, 1980b). Supralittoral wrack beds (especially Fucus) are clearly dominated by Lumbricillus lineatus, which was found in maximum densities of 1.7×105 ind.·m−2 on a North Sea beach (Giere, 1975) and about 75×104 ind.·m−2 on Baltic beaches (Backlund, 1945). Individual numbers of 50 to 100×104·m−2 should be the rule, which is comparable with tubificid densities in polluted freshwater areas (Brinkhurst & Kennedy, 1965). Intertidal mud flats harbour mainly naidids and tubificids (Table V). Due to an enhanced primary production of benthic microalgae by the emersion of wadden areas during low tide, the diatom-eater Amphichaeta sannio is

TABLE IV Some reports on maximum abundances of marine intertidal and supralittoral oligochaetes mostly belonging to the meiobenthos: IT, intertidal; SL, supralittoral

Species

Locality

Habitat

No. ×103·m−2

Reference

Aeolosomatidae Aeolosoma litorale

Sheltered detritus sands,

SL

18 Pfannkuche, 1980b

W.Baltic Tubificidae Aktedrilus monospermathecus

Exposed sands, N. Atlantic, Scotland

”

Detritus sands, North Sea SL

83.5 Pfannkuche, 1980b

”

Detritus sands, W. Baltic

SL

47.4 ”

Phallodrilus prostatus

Detritus sands, North Sea SL

101 ..

Enchytraeidae

IT

355 Giere, 1975

Interlinking of physical

708

Marionina subterranea

Exposed sands, N. Atlantic, Scotland

IT

1690 McIntyre & Murison, 1973

”

Exposed sands, N. Atlantic, Scotland

IT

5976 Giere, 1975

”

Detritus sands, North Sea SL

532 ”

M.spicula

Semi-protected sands, Brittany, France

IT

100 Lassèrre, 1967

”

Detritus sands

SL

556 Giere, 1975

M.subterranea and M. southerni

Detritus sands, W. Baltic

SL

M. subterranea and M. southerni

Protected sands, E. Baltic, SL Finland

39.9 Ax & Ax, 1970

Lumbricillus lineatus

Wrack beds (Fucus), North Sea

SL

1721 Giere, 1975

”

Semi-protected sands, Brittany, France

IT

900 Lassèrre, 1967

”

Wrack beds (Fucus), Central Baltic, Sweden

SL

750 Backlund, 1945

Enchytraeus albidus

Wrack beds, Central Baltic, Poland

SL

41 Pfannkuche, 1980b

60 Moszynski, 1930

temporarily the most dominant naidid in these biotopes. Up to 80×104 A.sannio·m−2 were counted in the intertidal mud flats of the Dollard Estuary (southern North Sea) during early spring (Koene, 1981). Heavy pollution probably favoured development of the rich populations (see p. 281) in the Forth Estuary (Table V; annual mean number up to 12×104·m−2 (Bagheri & McLusky, in press), but this species is also very frequent in shallow submersed beaches, as was demonstrated by Dahl (1960) and Arlt (1975). Paranais litoralis reaches high abundances in intertidal mud flats (Watling, 1975; Kendall, 1979), particularly if polluted (Bagheri & McLusky, in press), and in shallow bights (Arlt, 1975). Beside naidids, these areas are dominated by tubificids which represent an important standing stock throughout the year (Fig. 3), while naidids fluctuate considerably in seasonal abundance (see p. 193, Figs 2 and 5). Birtwell & Arthur (1980) found a maximum of 601×103 Tubifex costatus·m−2 in intertidal mud flats of the Thames Estuary (β-mesohalinicum). The annual mean of 439×103 ind.·m−2 clearly surpasses all other

TABLE V Some reports on maximum abundances of marine oligochaetes mainly from submersed or water-saturated areas: IT, intertidal; ST, subtidal

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Locality

Habitat

235

No. ×103·m−2

Reference

Megascolecidae Pontodrilus bermudensis Detritus sands, Indian Ocean, India

IT

50 Subba Rao & Ganapati, 1975

Naididae Amphichaeta sannio

Mud flats, Dollard Estuary North Sea

IT

800 Koene, 1981

”

Silty sands, Kattegat

ST

155.5 Dahl, 1960

”

Muddy bottom, Central Baltic

ST

45.1 Arlt, 1975

A.leydigii

Mesohaline ponds, North Sea

Paranais litoralis

Muddy bottom, Central Baltic

ST

”

Mud flats, Tees Estuary, North Sea

IT

”

Polyhaline mud flats, IT Forth Estuary, North Sea

”

Silty sands, Kattegat

ST

50 Muus, 1967

Nais elinguis

Detritus sands, Schlei, W.Baltic

ST

15 Pfannkuche, 1974

”

Sandy bottom, E.Baltic

ST

13.6 Laakso, 1969

Tubifex costatus

Mud flats, Thames Estuary, North Sea

IT

601 Birtwell & Arthur, 1980

”

Polyhaline mud flats, IT Forth Estuary, North Sea

”

Mud flats, Irish Sea

IT

”

Muddy bottom, Trave Mouth, W. Baltic

ST

Tubifex pseudogaster

Sandy bottom, Schlei, W.Baltic

ST

Peloscolex benedeni

Coarse sands, Channel

IT

”

Mud flats, Thames Estuary North Sea

IT

”

Polyhaline mud flats, IT Forth Estuary, North Sea

1200 Heip, 1971 73.1 Arlt, 1975 90 Kendall, 1979 156 Bagheri & McLusky, in press

Tubificidae

445 McLusky et al., 1980 40 Brinkhurst, 1964 63.2 Mein, 1979 35 Pfannkuche, 1980c 1000 Lassèrre, 1967 142 Hunter & Arthur, 1978 98 Bagheri & McLusky. in press

Interlinking of physical

710

”

Silty sands, Kattegat

ST

53 Muus, 1967

”

Coarse sands, Cape Cod, ST N. Atlantic

79.6 Cook, 1971

Monopylephorus indicus Brackish harbour mud, and M.waltairensis polluted by domestic sewage. India

ST

50–200 Subba Rao & Venkateswara Rao, 1980

Clitellio arenarius

Coarse sands, Channel

IT

150 Lassèrre, 1967

Tubificoides gabriellae

Muddy bottom, Chesapeake Bay, N. Atlantic

IT

Phallodrilus leukodermatus

Coralline sands, Bermuda ST Islands, Atlantic

10 Virnstein, 1976

≈100 Giere, unpubl.

Enchytraeidae Grania postclitellochaeta Coarse sands, Kiel Bight, ST W.Baltic

23.5 Pfannkuche, unpubl.

values which have been so far reported for macrobenthic marine tubificids. In the same estuary, Hunter & Arthur (1978) counted a maximum of 142×103 Peloscolex benedeni·m−2 from the β-mesohalinicum, while Bagheri & McLusky (in press) report peak values of 98×103·m−2 from polyhaline sites of the Forth Estuary (annual mean value 45×103·m−2). Lassèrre (1967) found even a maximum of 1×106·m−2 in intertidal coarse sands of the French Channel coast (Table V). The maximum densities quoted above refer mainly to the supralittoral hygropsammal (enchytraeids) and to eulittoral tidal flats (tubificids). Even in the shallow subtidal, however, many tubificids and few enchytraeids can attain remarkable abundance. In permanently submersed, often brackish fjords and bays, maximum densities of 40 to 80×103 ind.·m−2, mainly of Tubifex costatus, T.pseudogaster, and Peloscolex benedeni, are not unusual (Dahl, 1960; Muus, 1967; Cook, 1971; Erséus, 1976a; Pfannkuche, 1979, 1980c). In shallow subtidal bottoms (4 m in depth) in Bermuda, the gutless tubificid Phallodrilus leukodermatus has been found in ind.·m−2 in the upper 10cm of the calcareous sand (unpubl. data). The enchytraeid genus Grania is also often very abundant in sublittoral sands (Erséus, pers. comm.; pers. obs.). Pfannkuche (unpubl.) found a maximum of 23.5×103·m−2 G.postclitellochaeta in coarse sediments (10 m depth) of Eckernförde Bay (western Baltic). The eutrophicating effect of domestic sewage on some oligochaete species, adapted to low oxygen contents and formation of hydrogen sulphide, is apparent in an Indian brackish harbour basin where populations of the tubificids Monopylephorus indicus and M. waltairensis reached extreme densities (Table V; p. 282; Subba Rao & Venkateswara Rao, 1980). In general, tubificids and, less frequently, enchytraeids, can reach numerical importance down to ≈50 m depth. In deeper offshore bottoms, however, oligochaete population density is much reduced (Lassèrre, 1971 a; Erséus, 1976a), despite their rich species diversity (see Erséus, 1980a).

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BIOMASS When transforming abundance data into biomass values, on the basis of standard wet weights, one should be aware of the considerable discrepancies in weight within families, genera, and ecological groups (Table VI). Variable weights even between populations of one species have already been stressed by Giere (1975), and become particularly clear from data on Lumbricillus lineatus. Giere found 0.3 mg wet and ≈85 µg dry wt to represent relatively uniform figures for specimens from various habitats, whereas Dahl (1960) reported 1.5 mg wet wt for specimens from Danish shores. Regier (unpubl. data) recorded 147 µg dry wt for cultured animals which exceeds the above value by 80%. On the other hand, L. lineatus from Baltic beaches weighed only 60 µg dry wt. In a study of Bagheri & McLusky (in press), mean individual dry weight data varied from one sampling station to the other by 92% (from 0.095 mg to 0.182 mg) for Peloscolex benedeni and 60% (from 0.019 mg to 0.030 mg) for Paranais litoralis. When compared with 0.3 mg wet weight for this species (Bregnballe, 1961), the same value given by Dahl (1960) for the much smaller species Amphichaeta sannio seems very high. In connection with these discrepancies, individual weighing of oligochaete

TABLE VI Average wet weights (live weight in mg) of some marine oligochaetes

Species

Live weight

Reference

Naididae Amphichaeta sannio

0.3

Dahl, 1960

Paranais litoralis

0.3

Bregnballe, 1961

Nais elinguis

0.16

Ankar & Elmgren, 1976

Tubifex costatus

0.7 (immature)

Brinkhurst, 1964

”

2.3 (mature)

”

Tubificidae

”

1.5

Dahl, 1960

”

2.04

Birtwell & Arthur, 1980

T.pseudogaster

1.5

Pfannkuche, 1980c

Peloscolex benedeni

3.5

Dahl, 1960

”

3.5

Muus, 1967

Clitellio arenarius

1.6

Ankar & Elmgren, 1976

Tubificoides gabrlellae

1.8

Haven et al., 1977

Aktedrilus monospermathecus

0.017

Giere, 1975

Interlinking of physical

712

Enchytraeidae Enchytraeus albidus

≈10.0

O’Connor, 1967

”

4.1

Krüger, 1955

Lumbricillus lineatus

0.3

Giere, 1975

”

3.0

O’Connor, 1967

L.rivalis

9–10.0

”

L.viridis

13.0

”

Marionina subterranea

0.003

Giere, 1975

M.spicula

0.026

”

”

0.03

Lassèrre, 1969, 1971c, 1976

M.achaeta

0.03

”

samples after displacement of the gut contents seems much more useful than the calculation of biomass on the basis of standard wet weights. The conversion factors for calculation of dry weights from wet weights exhibit similar differences. Elmgren (in Ankar & Elmgren, 1976) gave a figure of 13% for Nais elinguis; Ankar (in Ankar & Elmgren, 1976) found the dry weight of Clitellio arenarius to be 18% of its wet weight. For the freshwater tubificids Limnodrilus hoffmeisteri and Tubifex tubifex 16% was reported by Brinkhurst (1970), and for enchytraeids, Giere (1975) calculated ≈30% (Table VII). Although reliable, ash-free dry weights of marine oligochaetes are rarely recorded. Judging from values of Giere (1975, Table VII), McLachlan (1977a), and Koene (1981) ranges of 0.1 to 1.6 µg could be representative for meiobenthic worms, and larger enchytraeids will attain 5 to 15 µg. The macrobenthic tubificids (e.g. Peloscolex benedeni) will weigh ≈100 to 250 µg ash-free dry wt (Gray, 1976). Despite their diversity depending on habitat and regional conditions, biomass values underline the importance of oligochaetes in some brackish-water and marine ecosystems. In comparison with other meiofaunal groups (Turbellaria, Nemertini, Nematoda, Harpacticoidea) oligochaetes represented in the wave-wash zone of sandy beaches in the Kiel Bight (western Baltic) one of the most important groups considering individual biomass within the length classes 0.5–1 mm, 1–3 mm, and 3–6 mm (Faubel, pers. comm.). Giere (1975) found the portion of meiobenthic enchytraeids to amount 20% of total

TABLE VII Weight data of some marine oligochaetes: average data for one individual based on several weighings of 100 individuals each, dry wt, after 80 °C for 24 h; ashfree wt, after 580 °C for 5 h (from Giere, 1975)

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Absolute data (µg) Species

Live wt

Lumbricillus lineatus

Percentage relation

Dry Ash-free Live wt wt wt

Dry Ash-free wt wt

306.0

83.9

5.0

100

27

1.7

26.3

9.8

0.6

100

37

2.3

M. subterranea (living, Sylt, Germany)

3.5

1.1

0.03

100

31

1.0

M.subterranea (preserved and stained from Loch Ewe, Scotland)

6.3

2.1

0.3

100

32

4.6

17.5

5.6

0.3

100

32

1.5

Marionina spicula

Aktedrilus monospermathecus

meiofauna at the high-water line at Firemore Bay (western Scotland). Nematodes clearly dominated in abundance, but in biomass the relation between oligochaetes and nematodes was 40:60. Arlt (1973) reported meiobenthic oligochaetes to exceed all other meiofauna in biomass in shallow bights of the Baltic. The biomass recordings of macrobenthic oligochaetes often exceed those from other faunal groups or even total macrofauna in various habitats (Table VIII): Tubifex costatus, 1201 g·m−2 (Birtwell & Arthur, 1980), 53 g·m−2 (Dahl, 1960); Peloscolex benedeni, 497 g·m−2 (calculated from data in Hunter & Arthur, 1978), 185g·m−2 (Muus, 1967). The maximum biomass of 2682 g·m−2 wet wt, equivalent to 725 g·m−2 dry wt, for Lumbricillus lineatus, aggregated in Fucus wrack beds of the North Sea (Giere & Hauschildt, 1979), represent an extraordinarily high value. In deeper waters, oligochaete biomass is clearly diminished. Ankar & Elmgren (1976) reported oligochaetes with a wet weight of only 0.57g·m−2 from the Askö-Landsort area (Baltic). The calculation of ‘biovolume’ serves as an additional means to illustrate population density, as shown by Giere (1975). A maximum biovolume: sediment volume ratio of 1:84 for Lumbricillus lineatus from natural habitats underlines the frequency of oligochaetes more clearly than do weight figures. This value is in the same range as the ratio of 1:75 for a population of Limnodrilus hoffmeisteri and Tubifex tubifex in a polluted creek (Giere, 1975, calculated after Brinkhurst & Kennedy, 1965). Moszynski (1930) opines a ratio of 1:100 as “optimal” for enchytraeids and values of about 1:1000 as “often occurring”. PRODUCTION Production data on marine oligochaetes are scarce due to difficulties in estimating standing stock data (see above), in assessing life cycles (no age-classes discernible) and metabolic activities. Their inaccuracy is enhanced as most of them are based on only few data derived from freshwater species (e.g. Ivlev, 1939; Teal, 1957; Johnson & Brinkhurst, 1971). A basic problem is the non-discriminating counting and weighing of “Oligochaeta” as the lowest taxonomic unit, which is common in many marine ecological studies.

TABLE VIII

Interlinking of physical

714

Maximum biomass (g live wt·m−2) of some oligochaete populations

Species

Habitat

Biomass

Reference

Tubifex costatus

Mud flat, β-mesohalinicum, Thames Estuary, North Sea

”

Soft bottom, α-mesohaline fjord, Kattegat

53 Dahl, 1960

Peloscolex benedeni

Soft bottom, α-mesohaline fjord, Kattegat

185 Muus, 1967

”

Coarse sands, subtidal, Cape Cod, N. Atlantic

278 Calculated from Cook, 1971

”

Mud flat, α-mesohalinicum. Thames Estuary, North Sea

497 Calculated from Hunter & Arthur. 1978

”

Mud flat, polyhalinicum, Forth Estuary, North Sea

85.2 Calculated from Bagheri & McLusky, in press

P. benedeni T.costatus

1201 Birtwell & Arthur, 1980

68.7 Pfannkuche, 1980c Sandy bottom, a-mesohalini-cum, Schlei-Förde, W. Baltic

T.pseudogaster Oligochaetes

Fine sand, β-mesohalinicum, N. Baltic

11.5 Ankar & Elmgren, 1976

Oligochaetes

Intertidal sands, N. Atlantic

43.9 Giere, 1975

Lumbricillus lineatus

Fucus wrack, supralittoral, North Sea

2682 Giere & Hauschildt, 1979

Amphichaeta sannio

Mud flat, polyhalinicum, Forth Estuary, North Sea

Paranais litoralis

”

33.0 ”

Dominant Oligochaetes

”

125.1 ”

6.9 Calculated from Bagheri & McLusky, in press

Aware of the fundamental differences in population dynamics between the families of marine oligochaetes, a resolution at least down to the family level seems indispensible. As shown in Figure 5, the seasonal abundance of naidids in a brackish-water fjord varied by more than 1300% of the annual mean. Also tubificid populations can fluctuate considerably in abundance (see Fig. 6). Quantitative data on mortality rates and the grazing effects on marine oligochaete populations are almost unknown. Growth rates, fertility, and reproductive cycles depend largely on the abiotic regime of the locality and its productivity (see p. 187 ff). Also the estimation of growth rates and metabolic activities (e.g. respiration) in laboratory cultures evokes many problems. Chua & Brinkhurst (1973) found the

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respiration rates of a mixed culture of T. tubifex, Limnodrilus hoffmeisteri, and Peloscolex multisetosus lower than in pure cultures while growth displayed a reverse relationship. (Similar difficulties could arise when transforming data from monospecific populations to conspecific ones). Ankar & Elmgren (1976) in their calculation of oligochaete production in the Askö-Landsort area used a turnover ratio (production/biomass : P/B) of 1.8, based on data of Johnson & Brinkhurst (1971) from Lake Ontario oligochaetes. These values would fit into the two-year life cycle assumed for Tubifex costatus and Peloscolex benedeni. Nevertheless, Johnson & Brinkhurst (1971) also reported P/B ratios of 12.3, 6.5, and 2.9 from the same region. High fission rates (p. 185) in Amphichaeta sannio can result in P/B ratios as extreme as 18 (daily production =20 mg C·m−2; Koene, 1981). The studies of Gray (1976), Warwick, Joint & Radford (1979) and Bagheri & McLusky (in press) base production of Peloscolex benedeni on a P/B value of 3 which Haka et al. (1974) had obtained for Finnish freshwater oligochaetes. Jonasson & Thorhauge (1976), studying the deep-water population of Potamothrix hammoniensis in Lake Esrom (Denmark) under various aspects, found P/B ratios of 1.35 to 0.78 in a four-year study (life cycle of P. hammoniensis is four years). They provided a new method for the estimation of production by using daily growth rates extrapolated to annual periods by calculation of the number of degree days available for growth. This method gave results comparable with the common method of Allen (1950). The figures given here underline the wide variability of so-called “constants”, apparently much depending on the local biome and the methods applied. ASSIMILATION Assimilation as the sum of production (0.78 g C·m−2·yr−1) and respiration (1.58 g C·m−2·yr−1) was determined by Warwick et al. (1979) for ≈9000 Peloscolex benedeni·m−2 (biomass 0.26 g C) to be 2.36 g C·m−2·yr−1. This paper quotes also consumption values of almost 8gC·m−2·yr−1. Total tubificid production was calculated here as 1.17 g C·m−2·yr−1. In the Forth Estuary, oligochaetes reached a production of 11.6 g dry wt·m−2·yr−1 (McLusky et al., 1980; Bagheri & McLusky, in press), which would correspond to ≈70 g wet wt (Table IX) using the conversion factor of 5.5 given for Clitellio arenarius and of 7.7 for Nais elinguis (Ankar & Elmgren, 1976). For freshwater tubificids, assimilation was determined by Johnson & Brinkhurst (1971) and Brinkhurst, Chua & Kaushik (1972) on the basis of laboratory measurements. Brinkhurst & Austin (1979) came to comparable results using a method based on the determination of the quantities of organic material present, left in the residue and voided as faeces. The need of individual measurements of the figures used for the calculation of production is also underlined by the great variability in recorded respiration rates even for one species (see the data of Dausend, 1931; Ivlev, 1939; Berg, Jonasson & Ockelmann, 1962; Aston, 1966; Palmer, 1968; Johnson & Brinkhurst, 1971; Brinkhurst et al., 1972). The uncritical acceptance of caloric equivalents from freshwater species and their transformation to marine species must also be considered rather inadequate. So far, almost all calculations of caloric equivalents are based on those of Ivlev (1939): Tubifex tubifex, 5547kcal·g−1 ash-free dry wt; Teal (1957): Limnodrilus hoffmeisteri, 760kcal·g−1 ash-free dry wt, and Cummins & Wuycheck (1971): Dero limosa, 5530 kcal·g−1 ash-free

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dry wt. Despite this cautious interpretation of the existing production values, the available estimates generally underline the importance of marine oligochaetes as producers in certain marine and brackish-water ecosystems (Table IX) which have been derived from abundance (Tables IV, V) and biomass

TABLE IX Annual production (g wet wt·m−2) and turnover ratio (P/B) for oligochaetes from some marine and brackish-water habitats

Species

Habitat

Production P/B

Reference

Meiobenthic oligochaetes

Silty sands, Greifswalder Bodden, S. Baltic

38.5

Meiobenthic oligochaetes

Exposed sands, Firemore Bay, N. Atlantic

40–70.0

Oligochaetes

Askö-Landsort Area, N. Baltic

Lumbricillus lineatus

Fucus wrack, Isle of Sylt, North Sea

400–650

Peloscolex benedeni

Intertidal mud flats, Forth Estuary, North Sea

60

3 Calculated from Bagheri & McLusky, in press

Amphichaeta sannio

”

2.4

3 ”

Paranais litoralis

”

7.8

3 ”

Dominant oligochaetes

”

≈70

3 ”

2.2

2 Arlt, 1973 3 Giere, 1975 3 Ankar & Elmgren, 1976 3 Giere, 1975

data (Table VI). So far, however, there is little evidence that these dense populations exert a similar meliorating effect on marine bottoms as has proved to be the case for enchytraeids in terrestrial soils (Nielsen, 1955c; O’Connor, 1967; Standen, 1973) and tubificids in freshwater bottoms (McCall & Fisher, 1980).

FOOD OF MARINE OLIGOCHAETES The early conception that aquatic oligochaetes are non-selective deposit-feeders or, in some cases, diatom-grazers, was derived from freshwater studies, is established in general textbooks and, although generally unexamined, is widely adopted for the marine species. Scrutinization of food demands in aquatic oligochaetes leads, however, to a far more differentiated view of food selection and preference. Again, there are few results specifically on marine forms, and one has often to extrapolate from limnetic studies.

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The trophic spectrum of marine oligochaetes can be grouped into five main categories. (1) Live microalgae (diatoms, flagellates). (2) Micro-organisms (bacteria, fungi, ciliates). (3) Particulate organic matter (detritus). (4) Fresh plant material (seaweeds) washed ashore. (5) Dissolved organic matter. The food uptake of most oligochaetes, however, overlaps these clearcut divisions which in nature are often difficult to separate. The muddy deposits, for instance, main food of large tubificids populating tidal flats and estuarine bottoms, contain, besides organic particles from decaying plants and animals (detritus), dense films of diatoms as well as rich bacteria colonies. On the other hand, ingesting algal thalli or Zostera leaves from wrack beds, enchytraeids which abundantly colonize the upper shore, not only devour plant cells but also bacterial slimes and diatoms (Fenchel, 1970; Harrison & Harrison, 1980), a mixture which renders a general categorization of these enchytraeids as ‘herbivores’ problematical. LIVE MICROALGAE The position of marine naidid species as primary consumers, selectively utilizing pennate diatoms, is relatively well and repeatedly documented (Bülow, 1957; Dahl, 1960; Pfannkuche, 1977; Ruyter van Steveninck, 1978). The gut of Amphichaeta sannio was found to be filled exclusively with diatoms; Koene (1981) believes this worm to select particular diatom species in the Dollard Estuary. The population cycle of Amphichaeta spp. in shallow muds correlated fairly well with blooming periods of these microalgae (Dahl, 1960; Heip, 1971; Pfannkuche, 1977, 1979; Ruyter van Steveninck, 1978; Koene 1981; see also p. 185). In Nais elinguis and Paranais litoralis, the intestine only occasionally contained diatoms indicating a reduced specialization for this diet (P.litoralis is held to be a general deposit-feeder, pers. obs.; Kendall, 1979). Pfannkuche’s (1977) and Moore’s (1978) results on freshwater oligochaetes suggest a larger trophic flexibility to be of adaptive value—diatom preference during blooming periods shifted to a diet of protozoans in summer (see below) and of detritus plus bacteria in winter. Occurrence of these naidids, specialized for epibenthic life in shallow brackish water, is in good correlation with the immense diatom production of mud flats richly provided with nutrients and light (several million cells·m−2 according to Fenchel & Straarup, 1971; Länge, 1980; Ramm, unpubl.). Hence, even a maximal consumption of well above 100 diatoms·day−1·worm−1 in Amphichaeta sannio, 200 cells in Nais and even 500 cells in the larger Paranais litoralis (Ruyter van Steveninck, 1978) which would mean a daily ingestion rate of about 50% their own body weight, causes no regulating effect by oligochaetes on the vast diatom stock. Even the maximum of 200–300 diatom cells·day−1·worm−1 reported to be taken by Amphichaeta sannio, whose population density can reach 80 ind.·cm−2 (Koene, 1981), would not significantly change this relation. According to Koene’s calculations, A. sannio could ingest as much as twice its own body weight per day when devouring the diatom Navicula salinarum. For interstitial

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tubificids and enchytraeids, populating preferably the subsurface layers of sandy shores, a trophic specialization on diatoms would be problematical because the abundance of these algae is much lower in sandy bottoms (Ramm, unpubl.), albeit the algae are known to live even at 20 cm depth (Pamatmat, 1968; Fenchel & Straarup, 1971). Hence, diatom skeletons attached to sand grains were only occasionally found in the gut of mesopsammic oligochaetes (e.g. Marionina subterranea) (Giere, 1975). MICRO-ORGANISMS Although there is no specific study available, scattered remarks in the literature, our own observations, and conclusions derived from other meiobenthic forms, render increasing evidence that bacteria form the basic trophic element of interstitial oligochaetes. The rich and diverse bacterial colonies attached to every sand grain (Meadows & Anderson, 1966, 1968; Steele, Munro & Giese, 1970; McIntyre & Murison, 1973; Alongi & Tietjen, 1980) provide an ample and spatially partitioned supply for bacteria-feeding oligochaetes even in ‘detritus-free’ exposed beaches. Consequently, in many psammobiotic oligochaetes, the gut contained only brownish, amorphous masses, probably of bacterial origin, as neither seaweed cells nor diatom skeletons could be detected (Giere, 1975). Parallel microscopical examination of ‘clean’ sand confirmed the availability of sufficient bacterial food; brown clusters of microbes covered the surface of the sand grains as a dense layer assimilating the rich nutrient content of interstitial water. The way oligochaetes detach and incorporate these micro-organisms is presumably more an “epipsammic browsing” than “sand-swallowing” which Boaden (1964) assumed to be the dominant mode of grazing in interstitial oligochaetes. According to our own studies on most various oligochaetes, whole sand grains are rarely found in the worms’ gut and the occasions are restricted to few species. Development of a muscular, reversible pharyngeal pad also indicates specialization for abrasion and suction of food items (Palka & Spaul, 1970). In contrast to the avoidance displayed by many (large) oligochaetes for fresh plant material (Stephenson, 1930; Backlund, 1945; Bülow, 1957; Giere, 1975; Hauschildt, 1978), the attractiveness of (aerobically) decaying seaweed is remarkable as seen from experiments with M.southerni, Aktedrilus monospermathecus (Jansson, 1968a), Lumbricillus lineatus (Giere, 1975; Hauschildt, 1978), and L.reynoldsoni (Tynen, 1969) in debris-inoculated sand. It becomes evident that bacterial activity plays a decisive rôle in the uptake of organic remains. An indication on the importance of bacteria not only for meiobenthic oligochaetes but even for voluminous macrobenthic marine tubificids might be the frequent aggregation of these worms around the redox potential-discontinuity layer of sediment cores (Pfannkuche, 1977, 1980d), a stratum in which bacteria are usually highly concentrated (Yingst & Rhoads, 1980). But also in the well oxygenated sand of the wave-wash zone, the inverse relationship between biomass of meiofauna (dominated by Lumbricillis lineatus) and bacteria (Meyer-Reil & Faubel, 1980) emphasizes the nutritional significance of these micro-organisms for marine oligochaetes. Ability to discriminate between grades of remineralization apparently occurs widely and has been reported especially for marine enchytraeids from wrack bed biocoenoses (Lumbricillus spp., Enchytraeus albidus; (Backlund, 1945; Tynen, 1969; Giere, 1971e; Schöne, 1971;

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Giere & Hauschildt, 1979). That this differentiating capacity is based mainly on the bacteria covering the thalli, was substantiated in two ways. First, both meiobenthic (Marionina) and macrobenthic (Lumbricillus, Enchytraeus) enchytraeids could be successfully reared on a bacterial diet developing on agar plates (Giere, 1975) without any detritus as substrate. Secondly, in a most interesting research project initiated by Brinkhurst and his coworkers in Toronto, it could be demonstrated that (limnetic) tubificids possess a differentiating digestion pattern of bacteria on detritus both regarding selective utilization by different worm species and selective assimilation of bacterial strains within the worms’ gut with a maximum bacterial reduction of 72%. In some species, even a meliorating effect of bacterial growth on organic remains after passage through the intestine is indicated (Brinkhurst & Chua, 1969; Wavre & Brinkhurst, 1971; Brinkhurst, Chua & Kaushik, 1972; Chua & Brinkhurst, 1973; Brinkhurst, 1974). Coler, Gunner & Zuckerman (1967) have shown common limnetic tubificids to avoid certain bacteria and to prefer others. Whitley & Seng (1976) specified the digestive selectivity of tubificids; in their experiments the worms digested only gram-negative bacteria enzymatically, whereas other microbial groups passed the intestine undamaged. Jansson’s (1968a) experiments which yielded differing results even with interstitial enchytraeids of one genus when Zostera leaves were offered, can possibly be interpreted in the light of correspondingly specialized demands in bacterial food. Although all four species tested were known from field studies to prefer the debris used, only two were attracted by the decaying plants. The negative response of the remaining two species may be due to the absence of the ‘right’ bacterial groups on the plant substratum. Limnetic and terrestrial enchytraeids, both small meiobenthic and voluminous forms, are also known to utilize bacteria as food (Kurt, 1961; Nielsen, 1961; O’Connor, 1967; Dash & Cragg, 1972). Whether bacterial biomass in the natural environment can suffice to maintain oligochaete populations under average field conditions remains to be solved. Nielsen (1961) calculated an amount of bacteria of 30 to 40 g dry wt·m−2·yr−1 as necessary to cover the food requirements of 50000 terrestrial enchytraeids·m2. In marine littoral sediments with often considerably more than 100 000 ind.·m−2, enchytraeids as well as naidids and tubificids (Gray, 1971; Lassèrre, 1971c; Arlt, 1975; Giere, 1975; Pfannkuche, 1980b), a corresponding food demand would result in about 80 g·m−2·yr−1 an amount which exceeds or is just within the range of bacterial productivity values from various marine sediments. In tidal muds, ZoBell & Feltham (1942) reported 39 g·m−2·yr−1; Gerlach (1978) calculated 104 g·m−2·yr−1 for the upper 5 cm of subtidal silty sand, assuming a bacterial P/B ratio of 21. Applying this figure for standing stock values of Baltic beaches (Meyer-Reil, Dawson, Liebezeit & Tiedge, 1978) one obtains 105 g·m−2·yr−1. Only in shallow silty substrata or in wrack beds does microbiol production occasionally lie above these values. On the basis of the above P/B ratio, the standing stock values of Dale (1974) would result in a production of ≈500g·m−2·yr−1. This could cover the energy budget of larger marine oligochaetes even considering food demands of other faunal elements. It should not be concealed that these figures are contradictory to the much smaller standing stock found by Rheinheimer (1977) and Meyer-Reil & Faubel (1980) for Baltic sandy shores and resulting in relatively low production rates. The nutritive rôle of other micro-organisms than bacteria, especially microfungi and protozoans for marine or brackish-water oligochaetes is insufficiently known. Judging

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from some studies on terrestrial enchytraeids, it is possible that selective discrimination is not restricted to bacteria but is also developed for microfungi (O’Connor, 1967; Dash & Cragg, 1972). Pfannkuche (1977) described the polyphagous Nais elinguis to shift seasonally from diatoms as food to flagellates (Euglena sp.) during spring and summer. Although ciliates have an important share in many food chains (Fenchel, 1967), their rôle as a direct source of food for oligochaetes seems to be quite restricted. Their abundance in mud might suffice occasionally to represent the main food item for Paranais litoralis (Dahl, 1960). On the other hand, Giere (1975) could not find any relation between number of ciliates and the rich occurrence of interstitial enchytraeids in Scottish beaches. Presumably, protozoans will, despite their rapid turnover rate, only represent an additional direct food source for marine oligochaetes. The recent studies of Briggs, Tenore & Hanson (1979) stress, however, the indirect relevance of ciliate activity as an important regulatory system for detritivores by fragmenting debris and stimulating bacterial growth by grazing. It must, as yet, remain undecided whether the engulfing of harpacticoids (nauplii?) by Nais communis, as evidenced through the most elegant and interesting immunological methods of Feller et al. (1979; Feller, pers. comm.) can be considered as food uptake. PARTICULATE ORGANIC MATTER Despite the potential selective uptake of bacteria and the successful culturing on purely microbial diets mentioned above, the main food ingested by oligochaetes in nature is organic matter which is particularly high in littoral sands and muds. Tubificids like Tubificoides (Peloscolex) spp. and Tubifex costatus, burrowing in the soft sediment of tidal flats, seem to ingest the mud non-selectively, but probably differentiate in digesting the various organic particles. In experiments with Limnodrilus hoffmeisteri, Poddubnaya (1961) observed cessation of feeding after all the organic content in the muddy material offered as food had been used up and merely its mineral particles remained. Wagner (1968), on the basis of experiments with freshwater tubificids, stressed the importance of finer sediment fractions (<63 µm in diameter) for oligochaete abundance and Juget (1979) underlined a relation between (small) particle size and gut diameter of (freshwater) oligochaetes, calculating a specific index for sediment consumability. For the enchytraeids common in the upper shore, a mixed diet (micro-organisms, debris and/or fresh plant cells) has been confirmed; Lumbricillus lineatus and Enchytraeus albidus not only take up but also utilize organic (plant) debris. Evidence from gut analysis (Backlund, 1945) has been experimentally corroborated by Hauschildt who reared Lumbricillus lineatus successfully in cultures with rotten Fucus thalli brought to almost completely sterile conditions by tetracycline hydrochloride (Giere & Hauschildt, 1979). This result with a modern bacterio- and fungistatic agent is in contrast to tests by Backlund (1945) who failed to maintain cultures of Enchytraeus albidus and Lumbricillus with “sterile” Fucus, which was dried, ground, and re-saturated with tap water. Shore enchytraeids are also occasionally attracted by decaying animals washed up by the waves (Stephenson, 1930). Michaelsen (1927) found E. albidus feeding on carion— like dead fish. Lumbricillus spp. can even regularly be encountered in the dense layers of

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disintegrating mud snails (Hydrobia) underneath algal wrack. Very little is known about the digestive processes in marine oligochaetes. Palka & Spaul (1970) studied Lumbricillus rivalis from sewage beds in more detail (their “L.lineatus” certainly is a wrong identification, see Tynen, 1969) and found the worms to assimilate preferentially the proteins from organic matter whose amount increases with aging due to micro-organisms growing on the surface (Fenchel, 1969, 1970; Heald, 1971). Palka & Spaul also noticed the ability for mutual conversion of carbohydrates, proteins and fats, the fats being stored in the chloragogenic tissue. They mentioned the worms’ inability to digest the incorporated cellulose. This corresponds with the results of Nielsen (1962) on terrestrial enchytraeids which needed bacterial cellulase to metabolize plant cellulose and partially explains the avoidance of fresh seaweeds by many littoral species. For marine enchytraeids in general, however, the validity of this symbiotic relation to bacteria needs further confirmation. In an interesting approach, Faubel & Meyer-Reil (1981) recorded the activity of various enzymes in L.lineatus from a beach of the western Baltic Sea by spectrophotometrical methods, a-amylase activity as a measure of enzymatic decomposition of carbohydrates thus may reflect digestive patterns and phases of food uptake in these worms. These authors monitored the highest activity in the morning, a gradual decline after 12.00 h with a minimum at midnight. Besides the usual temperaturerelated enhancement of enzyme activity (peak at 37 °C), they found an ecologically interesting relation to salinity. The Baltic worms showed maximum a-amylase activity in a brackish environment (salinity 8–24‰) which corresponds well to their home range of salinity changes. FRESH PLANT MATERIAL Whether or not bacteria act as intermediaries macerating plant cells, there is no doubt that many macrobenthic species, at least occasionally, forage also on fresh plant material. Our own field observations showed L.rivalis to take up small filamentous algae from sewage beds, and populations from the North Sea shore ingested fresh Enteromorpho thalli (see also Bülow, 1957). Seaweeds (Fucus vesiculosus, Ulva spp., and Cladophora sp.) along the Finnish coast were converted by Lumbricillus rivalis in laboratory cultures within a few days into a brownish amorphous mud. We also observed the eurytopic L.lineatus and Enchytraeus albidus to penetrate into the inner parts of Fucus and Zostera fronds, apparently feeding on the thin-walled internal cells (see also Backlund, 1945; Bülow, 1957; Tynen, 1969; Schöne, 1971) and soon changing the wrack into a slimy mass. Generally, there seems to exist a marked ability to differentiate between phytal matter (or the micro-organisms growing on them?); Fucus detritus is widely preferred, followed by aged Zostera (Bülow, 1957) and green algae like Enteromorpha. Ulva and Cladophora are less attractive and red algae as well as blue-greens are refused (Giere, 1971e, 1975). DISSOLVED ORGANIC MATTER Fairly recently it was ascertained that dissolved organic substances, absorbed across the body surface, are of relevance for oligochaetes. While Brinkhurst & Chua (1969) had

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already proved the direct absorption of solved amino acids through the gut without mediation of the gut microflora in the limnetic Tubificoides multisetosus, Siebers (1976), Siebers & Bulnheim (1977), and Siebers & Ehlers (1978) produced corresponding data for marine oligochaetes; Enchytraeus albidus can incorporate transintegumentally considerable amounts of amino acids and carbohydrates. This uptake depends on the salinity and Na+ -concentration in the ambient medium and can, in the case of glycine, cover up to 35% of the worm’s metabolic demands (Siebers & Bulnheim, 1977). Enrichment of neutral amino acids to a 9 to 15-fold increased concentration, relative to the surrounding water, proves this process to represent an active transport (Siebers, 1976). These figures on absorption, so far only experienced in one marine species, must be considered unexpectedly high for worms with completely normal oral food uptake. On the other hand, they help to explain the existence of mouth- and gutless marine tubificids, recently discovered in sublittoral coralline sands from Australian, Bermudian, and Caribbean reefs (Jamieson, 1977; Erséus, 1979b; Giere, 1979b). It is at present being studied (Giere) whether the glandular ‘warty’ cuticle, often found in these interstitial worms from several species and genera, has some bearing on their peculiar nutritive pathway in which bacteria are possibly involved. The extremely long body in relation to width, sometimes reported in these forms (Erséus, 1979b; Giere, 1979b) may well be interpreted as an effective means of surface enlargement. Beside the direct uptake route of dissolved compounds, the intermediate rôle of bacteria and their extracellular products gains increasing evidence in the food spectrum of marine oligochaetes. Bacterial slimes accumulate dissolved enzymes, bacteria produce mucopolysaccharide fibres and, thus, convert dissolved organic substances into particulate matter (Hobbie & Lee, 1980). This conversion process is also underlined by Meyer-Reil & Faubel (1980). They thought the intensive uptake of labelled glucose by Lumbricillus populations resulted from the grazing of “organic matter originally derived from bacteria (e.g. cells, slime layers)”. Bacterial uptake activity could then account for the rapid depletion of 3H-glucose in their tracer experiments. These results stress the ‘indirect route’ of incorporation of dissolved substances via particulate organic matter provided by micro-organisms as intermediaries. Whether ‘cultivation’ of discrete bacteria and algae by the rich mucous excretion of the worms themselves is relevant as an additional dietary facet for oligochaetes, as it is for nematodes (Riemann & Schrage, 1978; Warwick, 1980) is, as yet, not known. The rôle of intermodulating effects between single fractions of the food available to marine oligochaetes, emerging, for instance, from the activity of ciliates (p. 210), has been widely disregarded, although already Wavre & Brinkhurst (1971) and Chua & Brinkhurst (1973) underlined the interspecific action of three sympatric tubificid species mediated by differentiating uptake of their mutual faeces and selective digestion of the bacterial content. Certain unknown macro-molecular substances, sensed even from a distance, seem to increase assimilation of food and growth (Brinkhurst, 1980a), thus creating different trophic ‘niches’ and enabling coexistence of seemingly identical ecological types. In the marine biota, because tubificid species rarely live in huge ‘clusters’ as they do in some freshwater biotopes with almost 1 million specimens·m−2, displacing all other

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fauna, this enhanced exploitation of food reserves might here be of less importance than in polluted freshwater areas. In the strand line and underneath wrack beds, however, enchytraeids regularly occur in large populations (Giere, 1975). Here, corresponding interspecific regulation, be it by means of nutritional selectivity or chemical interrelations, could help to intepret the coexistence of ecologically similar species. Another important interacting process is exemplified by the grazing on diatoms which live on ‘aged’ detritus. Thereby, a wider surface area becomes available for bacterial settlement (Harrison & Harrison, 1980) which in turn can be rasped by bacteria-feeding meiobenthic oligochaetes. Similarly, the larger oligochaetes species could be supported by the grazing activity of general meiofauna on bacteria, which in the experiments of Ten ore, Tietjen & Lee (1977) led to an increased breakdown of plant debris and, eventually, to a richer supply of macrofauna with ‘aged’ detritus. Sediment bioturbation by oligochaetes and passage of sediment through their gut may also have a meliorating or ‘gardening’ effect on bacteria and, thus, be of benefit for bacteria-feeding meio- and macrobenthic oligochaetes (Gerlach, 1978; see above). Despite the scarcity of data referring particularly to marine species, this survey on the nutritive pattern of oligochaetes sets the trophic demands of this group on a substantially wider basis than considered earlier; far from being unspecialized, omnivorous ‘detritusfeeders’, the worms have developed specific preferences and specializations which connect them to several food chains (see Fig. 38, p. 292 ff). Hence, one is probably allowed to use the statement made by Reise (1979) for meiofauna also for oligochaetes— their “trophic links are just as diverse as taxonomic composition”. It is probable that this diversification is also crucial for the recent species extension in marine oligochaetes, as many of them live sympatrically in neritic sediments. Interspecific, often indirectly regulated trophic effects may explain the diversity of related species despite fairly identical habitat conditions. It also becomes evident that every consideration of oligochaete distribution and zonation has to take food availability and specialization into account as a partitioning factor (see p. 250 ff).

MARINE OLIGOCHAETES AS PREY FOR CARNIVORES The oligochaete body represents highly nutritious food; Cummins & Wuycheck (1971) ranked freshwater forms with 5575 cal·g dry wt−1 (average of several species), higher than polychaetes (3503 cal·g−1). Considering the large biomass oligochaete populations often develop in coastal habitats (see p. 201), they should be extensively utilized by predators, despite the generally smaller body size of marine species compared with their freshwater relatives (Diaz, 1980). Since Giere (1975) published a first compilation on oligochaetes as prey for marine fauna, more information has been accumulated in particular on the utilization of macrobenthic intertidal forms revealing the wide range of predator species, their selectivity in food demands, and changing proportion of oligochaete consumption. Most literature stresses that demersal young fish intensively and often selectively fill their gut with oligochaetes. Müller (1968) found in western Baltic shallow flats the brackish water naidid Paranais litoralis to be the preferred food for young flounders and

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plaice. Although there were strong seasonal and diurnal fluctuations in the food of these fish due to a changing selectivity, he ranked the oligochaetes with polychaetes as the predominant food of the 30–35 mm long juvenile flat fish with up to 100 worms per stomach examined. These figures principally agree with Muus (1967) and Arndt & Nehls (1964). The latter found, besides P.litoralis, the enchytraeid Lumbricillus lineatus to be 53% of the food of these fish and reported a maximum stomach content of even more than 200 worms. Similar, but somewhat vague data on “oligochaetes” or “enchytraeids” as important food for young brackish-water fish are given by Ankar (1977), Elmgren (1978), and in other studies from the Baltic (Rauschenplatt, 1901; Hass, 1939; Smidt, 1951). One of the most specific sets of data, however, published by Bregnballe (1961), again characterizes Paranais litoralis from flats in the Aarhus area (Kattegat, salinity 18– 24‰) to be the most important food for plaice and flounder of the O-group. It was calculated that at least 50% of the Paranais populations are consumed by these fish which devoured 250 worms·fish−1·day−1equivalent to 75 mg wet wt; the oligochaete stocks are thus being strongly regulated by predation. Similarly, detailed results on heavy grazing activity of flat fish on the common marine tubificid Peloscolex benedeni are given by Van den Broek (1978) who studied stomach contents of fish from the Medway Estuary (U.K.). Up to 67% of flounder and plaice stomachs contained remains of this oligochaete. Although changing seasonally (maximum in early spring, minimum in late summer) and depending on age class (maximal uptake, 38–50%, in the 2-yr age class, less in the O-group), P. benedeni, on average found in 40.5% of flounder and 31% of plaice stomachs, was almost always the dominant food item of these economically important fish. Two facts indicate that P.benedeni was selectively picked up from the substratum: Tubifex costatus, similarly common in the area, was rarely found in the stomachs, and maxima of Peloscolex consumption did not coincide with oligochaete abundance. Summers (1980) reports corresponding fluctuations in the consumption of oligochaetes by flounders sifting the worms out of the mud from the middle region of the Ythan Estuary (Scotland). P.benedeni amounted to 34% of the food in May/June. Particularly in the upper estuary, ‘other oligochaetes’ represented 26% of the fish’s total food in June. There is scant information of other fish species feeding on oligochaetes; this scarcity of data may probably be attributed to the only moderate attention given to species not exploited by fisheries and, thus, does not necessarily reflect biological conditions. Young gobiids from the Schlei-Förde (western Baltic) are recorded by Westphal (1979) as devouring preferably Paranais litoralis. While he ranks this naidid, together with Nais elinguis, to contribute 75% of the food for Pomatoschistus spp., tubificids like Peloscolex benedeni and Tubifex pseudogaster, although amply present, contribute only 15–20% and the enchytraeid Lumbricillus only 3%. Altogether, in summer months, oligochaetes provide 50% of the fish’s food. For young whiting (Merlangius merlangus) and pouting (Trisopterus luscus), both Peloscolex benedeni and Tubifex costatus were of minimal nutritional importance in an English estuary although they were common in occurrence (Van den Broek, 1978). Ankar (1977) reported Gobiidae preying on oligochaetes in the northern Baltic. Reise (1977) found the gobiid, Pomatoschistus microps, in cage experiments on North Sea mud flats to deplete Peloscolex benedeni. This is in contrast to the comparable experiments by Virnstein (1977) from American shallow estuarine

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bottoms in Chesapeake Bay, where Tubiftcoides gabriellae, although numerically dominant, was largely unaffected by any predation by fish and crabs. Berge & Hesthagen (1981) also reported from the Oslo Fjord that population density of “oligochaetes” was not significantly changed by the grazing of Pomatoschistus microps, although representing the most important food component. Similarly, Schmidt-Moser & Westphal (1981) found only little evidence for a grazing effect on the naidid population in the Schlei-Förde, although these epibenthic worms were selectively preferred by P.microps. That enchytraeid populations (mainly Lumbricillus spp.) are also excellent food for young sole, emerged from British culture experiments where local seaweeds are rinsed for obtaining the worms (M.Fonds, pers. comm.). It is interesting to note that studies from the Baltic shores provide by far the bulk of information on oligochaetes as fish food. One may speculate whether this reflects the general accommodation of most marine oligochaetes to brackish water conditions. The second macrobenthic group reported to regularly ingest marine oligochaetes are Polychaeta. According to Rees (1940) and Yablonskaya (1953), especially the omnivorous and eurytopic Nereis diversicolor, often populating flats in 1000 ind.·m−2 can reduce oligochaete numbers (probably mainly naidids and Peloscolex benedeni) so heavily that they are inversely distributed in time and space. In laboratory cultures, Nereis diversicolor, and also N.virens, readily took Enchytraeus albidus (Ivleva, 1970; Goerke, 1971). In the eastern Baltic, the carnivorous polychaete Harmothoe sarsi was reported by Sarvala (1971) to include meiobenthic oligochaetes into its diet (see also Fig. 9 in Elmgren, 1978). There is good evidence (Plagmann, 1939; Muus, 1967; Scherer & Reise, 1981) from Baltic and North Sea flats for the predacious activity of small crustaceans like young shore crabs (Carcinus maenas) and shrimps (Crangon crangon) on oligochaetes such as Paranais litoralis, most available due to its epibenthic life, Peloscolex benedeni, and Tubifex spp. Reise (1977) found Peloscolex benedeni to be the dominant prey of these crustaceans and young gobies in cage experiments. Compared with the caged population he recorded an increase of up to 23 times in the uncaged stock. Depletion depending much on fluctuations in the predator population was most drastic from June to October. Interestingly enough, P.benedeni and Tubifex sp. were preferred by young female shore crabs only (Scherer & Reise, 1981). In contrast to Virnstein’s (1977) results, who could not find marked changes in abundance of oligochaetes after excluding predation, Bell & Coull (1978) reported feeding of the grass shrimp (Palaemonetes pugio) on Nais communis in experimental “salt marsh tanks”. Feller (pers. comm.) gave immunological proof for the presence of Nais communis tissue in the intestine of the American mud-snail Ilyanassa obsoleta from Spartina marshes. Even if this uptake were non-selective just by moving around in the superficial mud, stirring up the sediment with their mouth parts, and picking out digestible particles, the predatory effect on oligochaetes was acute. Ingestion of oligochaetes by amphipods like Pontoporeia spp. (Elmgren, 1976) seems to be similarly accidental during the process of non-selective uptake of deposits. The predacious priapulid Halicryptus spinulosus from the eastern Baltic was reported by Ankar (1977) regularly to consume Peloscolex benedeni and “naidids” among other annelid worms. There is some evidence on the extent and selectivity of nemerteans foraging on littoral

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oligochaetes. Whereas most of the voracious species probably engulf oligochaetes nonselectively (Jennings & Gibson, 1969) beside other prey available, Lineus ruber, a common littoral species, lived during summer predominantly on Clitellio arenarius (Jennings, 1960). Lineus ruber can be readily fed with oligochaetes in experiments (Bartsch, 1975). In suitable habitats, population densities of nemerteans can be remarkably high for a predacious group (Muus, 1967), so they may exert a substantial reduction in oligochaete biomass but, again, detailed pertinent studies are lacking. A negative impact chironomid larvae can exert on larger tubificids, resulting even in “competitive exclusion”, was suggested by Brinkhurst & Kennedy (1965) for freshwater species and confirmed by Loden (1974) who found after gut analysis several chironomid species devouring various oligochaetes. In coastal areas, this may be of relevance only in the oligohaline zone of estuaries and in brackish basins like the eastern Baltic Sea. The same may also apply to the hirudinid Erbopdella octoculata which eagerly and preferably forages on oligochaetes (Pfannkuche, Jelinek & Hartwig, 1975) and can proceed into areas of low salinity. The semi-terrestrial wrack beds, apparently, are a zone of intensive and selective feeding on the abundant enchytraeids Enchytraeus albidus and Lumbricillus spp. Backlund (1945) in his valuable monograph stated that the araneid Erigone longipalpis and many carnivorous beetles (Hydrophilidae, Staphylinidae) and their voracious larvae prey regularly on these worms. There is surprisingly little reliable evidence on water birds utilizing the rich and nutritious food source that oligochaetes represent especially in the upper tidal flats and strand line. General assumptions were based originally on observation of waders probing and pecking in wrack and mud, but recent analyses of stomach contents have proved that birds consume considerable quantities of marine oligochaetes (probably tubificids) in the tidal flats of Teesmouth (Evans, Herdson, Knights & Pienkowski, 1979; also Bryant, pers. comm.). Redshank (Tringa totanus) and shelduck (Tadorna tadorna) fed chiefly in areas rich in “small oligochaetes”. Whereas for these birds as well as for knot (Calidris canutus) oligochaetes were only a small part of their forage, 65% of the gut content in dunlins (C.alpina) consisted of these worms (judging from Gray, 1976, mainly Peloscolex benedeni, Tubifex pseudogaster, Paranais litoralis). Warnes (1981), working in the Forth Estuary, also found that Peloscolex benedeni is a significant component in the diet of T.tadorna in particular. These results underline the importance of oligochaetes as bird food, at least in favourable habitats, and also indicate a selective uptake since their natural abundance in the Teesmouth flats (Gray, 1976) was less than the figures from gut contents suggest. In polluted limnetic bottoms, tubificids (Tubifex tubifex, Limnodrilus hoffmeisteri) can also play a dominant rôle (31%) as food for ducks, as analysed by Rofritz (1977). He speculated that the scanty number of reports on oligochaetes as bird forage may result from the rapid digestion rate with which this highly nutritious and softbodied food passes the avian intestine and so tends to become unidentifiable or overlooked. Turning now to meiobenthic predators feeding mainly on smaller oligochaetes, attention is drawn to the Turbellaria. Their major importance is pointed out by the variety and eurytopic occurrence of predacious species reported to attack oligochaetes. So many forms belonging to various families and living in geographic-physiographically differing

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habitats have been described as devouring or sucking out oligochaetes, that Bilio (1967) ranks them as the main food at least for species from the North Sea salt marshes. Observations indicate that many turbellarians feed selectively and predominantly on oligochaetes (often even on relatively large species like Lumbricillus lineatus), viz. Coelogynopora schulzii, Uteriporus vulgaris (Proseriata) (Bilio, 1964, 1966, 1967), Proxenetes spp. (Typhloplanoida) (Bilio, 1965), Pseudograffila arenicola (Dalyelloidea) (Pfannkuche, 1977). Oligochaetes as prey besides other meiofauna and diatoms are mentioned for a wide range of turbellarians, from Otoplanida (Sopott, 1973) devouring small interstitial worms in beaches, to Kalyptorhynchia (Bilio, 1965) attacking even large worms in brackish muds. Corresponding accounts from various seas and biotopes stress the considerable rôle marine Oligochaeta play as food for turbellarians (Baltic: Luther, 1960; Straarup, 1970; North Sea: Den Hartog, 1964; American Atlantic Coast: Riser, pers. comm; Atlantic coast of Brittany: pers. obs.). Lumbricillus rivalis serves as rich food in sewage beds for freshwater Tricladida (Reynoldson & Seften, 1976; Sigurjonsdottir & Reynoldson, 1977). Dörjes (1968), investigating a North Sea beach, ranked the impact of these predators so high that he explained a mutual displacement between interstitial oligochaetes and turbellarians by predative grazing. This is in some contrast to Bilio’s (1965) results from North Sea salt marshes. While Bilio could not find a direct relation between turbellarians and oligochaetes, he pointed out that the average numbers of oligochaetes were unusually high in samples with rich turbellarian populations but did not give any explanation for this. Another group of meiobenthic predators, the Acarida, seem to exert some predative pressure on oligochaetes. Several semi-terrestrial and marine Gamasida and Halacarida have been observed to suck out marine enchytraeids and naidids (Lohmann, 1893; Schuster, 1962; Kirchner, 1969). The recent and most detailed article on coastal mites by Bartsch (1974) especially justifies the ranking of some forms like the common Halacarellus as oligochaete-preferring animals. It is probable that carnivorous nematodes, common in habitats rich in oligochaetes, use this food source but as yet there is no definite confirmation from field studies. Despite the increased number of studies mentioning oligochaetes as food of marine predators, most data only descriptively report ingestion of these annelids without any quantification. This lack of detailed information renders it impossible as yet to substantiate the nutritive rôle of oligochaetes by consumption rates or even mortality calculations. The wide array of oligochaete-grazing forms, reflecting the large range of marine oligochaetes which overlaps the macro-meiofaunal size thresholds, can be put into four major macrofaunal groups (young fish, polychaetes, crustaceans, and shore birds) and two meiobenthic categories extensively utilizing oligochaetes (Turbellaria and Nemertini). All other forms mentioned here for completeness sake, probably are of subordinate range (see Fig. 38, p. 294). In general, oligochaetes as a trophic component are of less relevance in the marine biota, where they are widely substituted by polychaetes, than in limnetic habitats, where they represent the bulk of the ‘worms’ (Pfannkuche, 1977; Diaz, 1980). Consequently, in bordering zones like estuaries, carnivor-ous feeding on oligochaetes is held to be of intermediate importance, i.e. is enhanced relative to marine biotopes, but reduced

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compared with limnetic habitats. In extreme marine habitat conditions like polluted coastal areas, oligochaetes, however, often re-gain considerable trophic relevance (Diaz, 1980). Another reason for the more reduced rôle of marine oligochaetes as food for higher members of the food chain may be, according to this author, their attenuated susceptibility to predators because the marine tubificids do not ‘wave’ with their caudal ends above the sediment surface as do their limnetic relatives. This reduced activity will make them less detectable. Despite all these restrictions, Diaz’s (1980) conclusion may well be justified that in coastal areas, particularly in estuaries and lagoons, marine oligochaetes are most important as food for macrofauna. The tentative character of these results, indicates that further evidence is urgently needed.

MARINE OLIGOCHAETES AS HOSTS FOR PARASITES Marine Oligochaeta are, as are their limnetic and terrestrial relatives, commonly hosts of protozoan parasites which almost exclusively belong to the orders Astomata (Ciliata: Holotricha) and Gregarinida (Sporozoa: Telosporidia). The numerous records of these parasites focus mainly on the morphology of the protozoans; the older papers particularly are often incomplete and superficial in their description and controversial in their nomenclature. Even the few more detailed reports containing thorough and modern morphological analyses, complete developmental cycles, and careful taxonomic reasoning, usually describe the parasitization from the protozoological points of view. This present review will stress the oligochaetous aspects without dealing too much with protozoan structural and taxonomical problems. Hence, the tables on parasites (Tables X and XI) quote the protozoans with their original names and do not enter into the general discussion on protozoan synonymons. Most people studying oligochaetes

TABLE X Astomatous ciliate parasites in marine oligochaetes

Host

Parasite

Area

Reference

Clitellio arenarius

“Opalina” (Anoplophrya lineata)

Hebrides Islands

Claparède, 1861

”

“Opalina” filum

Great Britain

Lankaster, 1870

”

”

Seashore, Isle Tatihou, English Delphy, 1922a Channel

”

Anoplophrya filum

Seashore, English Channel

Delphy, 1922b

”

A.fusiformis

”

”

”

A.debaisieuxi

”

”

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”

A.filum

Seashore, Roscoff, English Channel

Dehorne, 1927

”

Radiophrya grandis

Murman shores

Frolova, 1957

Host

Parasite

Area

Reference

Phallodrilus prostatus

Anoplophrya sp.

Brittany coast

Lassèrre, 1967

”

A. nodulata

Baltic and North Sea shores

Hartwig & Jelinek, 1974

”

”

Schlei-Förde, W.Baltic

Pfannkuche, 1977

Aktedrilus monospermathecus

”

”

Hartwig & Jelinek, 1974

Spiridion insigne

”

”

Pfannkuche, 1977

Lumbricillus lineatus

A. polydorae

”

Hartwig & Jelinek, 1974

”

“Opalina” (=Anoplophrya) pachydrili

Hebrides Islands

Claparède, 1861

”

Anoplophrya pachydrili

English Channel

Delphy, 1922a,b

”

A. elongata (in coelomic cavity!)

”

”

”

A. fusiformis

”

”

”

A. filum

Murman shores

Frolova, 1957

”

A. polydorae

Schlei-Förde, W.Baltic

Pfannkuche, 1977

L. “lineatus” (=rivalis)

Radiophrya pachydrili

Inland waters, Germany

Meier, 1954

”

Mesniliella elongata

”

”

L. rivalis

Radiophrya prolifera

Murman shores

Frolova, 1957

”

Anoplophryinae

Sewage beds, U.K. Reynoldson, 1939b

”

Anoplophrya filum

Schlei-Förde, W.Baltic

Hartwig & Jelinek, 1974; Pfannkuche, 1977

Lumbricillus sp.

Anoplophryinae

Spitsbergen, Bear Is.

Stephenson, 1925

Enchytraeus

”

”

”

Enchytraeus albidus

Anoplophrya filum

Schlei-Förde,

Hartwig &

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W.Baltic

Jelinek, 1974

”

Anoplophrya sp.

Soil, U.K.

Stirrup, 1913

”

Mesniliella fastigata

”

”

”

”

Inland waters, Germany

Meier, 1954

”

”

Kiel Bay, Baltic

Möbius, 1888

”

”

Inland waters, E.Europe

Heidenreich, 1935

”

”

Murman shores

Frolova, 1957

”

Radiophrya prolifera

Inland waters, E.Europe

Heidenreich, 1935

”

”

Murman shores

Frolova, 1957

”

”

Inland waters, Germany

Meier, 1954

Marionina welchi

Mesniliella sp.

Atlantic coast, U.S.A.

Giere, unpubl.

M. crassa

Radiophrya prolifera

Murman shores

Frolova, 1957

”

Anoplophrya filum

”

”

Marionina sp. (juveniles)

A.nodulata

Schlei-Förde, W. Baltic

Hartwig & Jelinek, 1974

”

A. polydorae

”

”

Marionina southerni

”

”

Pfannkuche, 1977

Cernosvitoviella immota Mesniliella sp.

Baltic shores, Germany

Giere, unpubl.

Paranais litoralis

Ciliata, unidentified

Kattegat region

Dahl, 1960

P. “elongata”

Anoplophrya paranaidis

Italian shores

Pierantoni, 1909

have frequently observed parasitized worms without mentioning or even describing them, the lists in Tables X and XI cannot claim, therefore, to be complete. Often large astomatous ciliates of the genera Anoplophrya (without endoskeleton), Mesniliella (with endoskeletal spine; see Fig. 8) and Radiophrya (with V-shaped cytoskeletal rod) are found to live in the gut of enchytraeids, tubificids, and naidids. Most species move freely in the gut cavity, others possess simple organs for loose attachment to the inner walls of the intestine or foregut, whence they absorb dissolved nutrients without apparent pathogenic effects on their hosts despite the often very massive infection (Stephenson, 1925, 1930). According to Dahl (1960), the heavy infestation of Paranais litoralis with ciliates, which made the posterior end of the gut “swollen with infusoria” and caused an unusual thickening of the chloragogenic layer, “may well be an important reason” for the population minimum in August. This statement should,

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however, be treated with caution as similar breakdowns of the population in late summer are reported without the worms being infested (see p. 195). Reporting on Anoplophryinae densely packed in the intestine of littoral enchytraeids, Stephenson (1925) found parasitization in the polar shores (Spitsbergen, Bear Island) to go far beyond what was common in British areas. Astomata seem to represent a transition between parasites and commensals. Multiplication by cross-division through the formation of ‘budding chains’ (Fig. 9) is common. Transfer to a fresh host by encystment, certainly of advantage in semiterrestrial shore species, has not yet been observed. Corresponding to their rather loose relation to the host’s body, Dahl (1960) mentions even frequent escape of intestinal ciliates through the anus of naidids, host-specificity does not seem to be very strict; the same ciliate species often inhabits several not even closely related oligochaete species (Table X; Frolova, 1957; Hartwig & Jelinek, 1974). Among the gregarinid sporozoans, the second group of common oligochaete parasites, the Monocystidae seem to represent the only family of relevance for marine species. Scanning through literature, no reference could be found to any marine tubificid infected by these parasites (Table XI); this probably indicates more a lack of information than a host-specific restriction since tubificids from limnetic habitats are reported to contain gregarines. Critical revision of the various forms parasitizing the body cavity or seminal vesicles (Giere, 1971a) and listed in Table XI, renders it probable that the different genera can be duly regarded as members of the large genus Monocystis. Often the figures are too sketchy and the descriptions too scant for one to establish new genera with any reliability. The same reasons make if difficult to synonymize the two intestinal gregarines (Kölliker, 1848; Claparède, 1861), certainly not belonging to Monocystis (Table XI). Few papers describe all taxonomically important stages of the developmental cycle, especially the shape and size of sporocysts and the morphological details of adult gregarines and syzygies. The more ‘internal position’ of the infectious stages (coelomic cavity, seminal vesicles) and their complicated developmental route through the host’s body have resulted in a fairly close adjustment of the parasitic cycle to the host’s physiology, growth, and maturation time (Giere, 1971a,b.). The young trophozoites which often cause hypertrophic growth of the worm’s chloragocytes were predominantly

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Fig. 8.—Two astomatous ciliates (Mesniliella?) in the intestine of the enchytraeid Cernosvitoviella immota (ZEISS interference-contrast, microflash photograph from living specimens).

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Fig. 9.—Budding chain of an astomatous ciliate (Mesniliella?) in the intestine of the enchytraeid Cernpsvitoviella immota (ZEISS interfyrencecontrast, microflash photograph from living specimen).

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Fig. 10.—Dense infection with sporocysts of Monocystis rhabdota in the posterior coelomic cavity of Lumbricillus lineatus (ZEISS interference-contrast, microflash photograph from living specimens) (from Giere, 1971b).

TABLE XI Gregarines in marine oligochaetes

Host

Parasite

Site of infection

Area

Reference

Lumbricillus lineatus “Gregarines”

Coelomic cavity

Clyde area, North Sea

Stephenson, 1913

”

Gregarina saenuridis

”

Sea shore, Channel

Delphy, 1922a

”

Monocyslis lumbricilli

Seminal vesicles

Sylt, North Sea

Giere, 1971a

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”

M. rhabdota

Coelomic cavity

”

Giere, 1971b

”

Monocystis sp.

Chloragocytes

Helgoland, North Sea

Giere, unpubl.

L. Oligochaetocystis “lineatus” (=rivalis) pachydrili

Coelomic cavity

Inland waters. Meier, 1956 Germany

L. rivalis

Gonospora pachydrili

Seminal vesicles

Inland waters, Vejdovsky, Europe 1889

”

Monocystis sp.

Coelomic cavity

Sewage beds, U.K.

Reynoldson, 1939b

L. viridis

Monocystis sp.

Chloragocytes

Helgoland, North Sea

Giere, unpubl.

Lumbricillus sp.

Monocystis sp.

Coelomic cavity

Schlei-Förde, W. Baltic

Pfannkuche, 1977

”

“gregarines”

Coelomic cavity, intestinal walls, gut lumen?

Spitsbergen, Bear Is.

Stephenson, 1925

Enchytraeus sp.

”

”

”

”

Enchylraeus sp.

Monocystis sp.

Coelomic cavity

Schlei-Förde, W. Baltic

Pfannkuche, 1977

E. albidus

“Gregarina” enchytraei

Coelomic cavity, seminal vesicles

Soil, Russia

Radkewitsch, 1869

Pontodrilus bermudensis

Monocystis pontodrili

Coelomic cavity

Brackish shores, India

Subba Rao et al., 1976

Lumbricillus semifuscus

aff. “Monocystis” sp.

Intestine

Hebrides Islands

Claparède, 1861

Intestine

not mentioned Kölliker, 1848

Enchytraeus albidus “Gregarina” enchytraei

encountered in juvenile Lumbricillus lineatus, whereas adult gregarines, syzygies, and gametocysts were most numerous in fully mature worms. In submature oligochaetes, however, all developmental stages of the monocystids were found which demonstrates that synchronization is neither strict nor even obligatory. The fact that the conformity of parasitic stages was always highest in the seminal vesicles and far less expressed in the coelomic cavity of the (posterior) body indicates some adjusting physiological influence possible originating from maturation processes. This is emphasized by the phenomenon that even in immature L. lineatus, it was the future region of the yet undeveloped genital organs which was predominantly parasitized (Giere, 1971a). In general, the body cavity anterior to the seminal vesicles was always free of monocystids and the posterior end of the body less infected than its central region containing the genitalia. An interrelation between worm and parasite was also expressed by the monospecific infection of one oligochaete population (L. lineatus) by just one monocystid. Separate

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populations of the same host species were found to be parasitized by different gregarines, which renders specificity in parasite-host relations rather complex. Infection of oligochaetes like L. lineatus and Enchytraeus albidus with parasites is often 100% and sometimes the worms’ coelomic cavity literally seems to be packed with sporocysts (Fig. 10; see also Delphy, 1922a); the worms’ activity, resistance, or productivity does not, however, suffer serious aggravation. We have never observed histological malformations affecting larger regions nor any sterilization effects despite this heavy parasitization. Stephenson (1925) did not exclude a relation between degenerative phenomena in the intestinal epithelium and the massive occurrence of gregarines; but he also found this degeneration of specimens without evident parasitization. This has been misinterpreted by Reynoldson (1939b) who quoted Stephenson’s cautious observations as intestinal “wall degeneration” and “blocking” of the lumen by ciliates (!). Later, this was unduly exaggerated by Hartwig & Jelinek (1974) who, in turn, quoted Reynoldson (1939b) to have observed “clogging” and subsequent “degeneration of the intestinal tube” caused by ciliates, although Reynoldson expressively stated “no pathogenic conditions” to be evident. Parallel maturation periods of oligochaetes and parasites (richest occurrence of mature gregarines and ripe gametocysts), a phenomenon well known in lumbricids, seems to a certain degree to exist also in E. albidus and Lumbricillus spp. Despite a persistent infection through the year (Kölliker, 1848; Delphy, 1922a), Giere (1971a,b) found high percentages of mature L. lineatus from the North Sea shores in the warmer season with particularly large numbers of syzygies, gameto- and sporocysts. Whether the numerous degenerative and aberrant stages, frequently observed among the monocystids, result from some defensive reaction of the worm remains uncertain. Stephenson (1930) describes gregarine cysts to degenerate possibly by the action of numerous phagocytes, but the very resistant sporozoites to remain alive even when incorporated by phagocytes. Mature trophozoites seem to remain ‘unattacked’. The high rate of infection and the eurytopic nature of the hosts along our shores are in strange contrast to the rather local occurrence of parasitized populations. In L. lineatus, only some hundred metres away from areas without any infection, populations were found 100% parasitized by monocystids (Giere, 1971a,b). Comparatively few other parasites like Actinomyxidia are found in marine oligochaetes: Hexactinomyxon psammoryctis (Actinomyxidia) in Clitellio arenarius (Caullery & Mesnil, 1904); the rotifer Albertia naidis in the intestine and coelomic cavity of Nais elinguis and N. variabilis (Erséus, 1976b; Pfannkuche, 1977). Cestodes of the genera Archigetes and Caryophyllaeus, regularly encountered as parasites in limnetic tubificids, have not yet been described in the intestine of marine or brackish-water species, nor could we find any report of nematode parasitization in marine oligochaetes.

HABITAT CONDITIONS AND OLIGOCHAETE OCCURRENCE ABIOTIC FACTORS “Inhomogeneity” and “patchiness” are typical terms in reports characteriz-ing marine

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oligochaete distribution. These terms are concomitant with the unstable habitat conditions in intertidal and shallow subtidal reaches of the sea, where heavily fluctuating environmental factors intermingle and create a complicated and interdependent web of ecofactors in which the oligochaetes have to orientate themselves. Often provided with unusually wide tolerance ranges (see p. 259), many survive even extreme physiographical conditions,

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Fig. 11.—Transect through a Danish beach showing gradients of some

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environmental factors and the distribution of three interstitial oligochaetes (modified after Fenchel, 1978).

but they respond nevertheless to minor environmental oscillations and maintain themselves in more restricted preference zones which results in a complex and changing distribution pattern difficult and often even impossible to disentangle and to relate to single determinants. Figure 11 illustrates the divergent distribution of several interstitial oligochaete species in a Danish beach as an example of the animals’ response to the array of environmental conditions. The intermingling action of ecofactors in the field will inevitably lead to some overlap in discussing the single operative features. Moreover, without experiments (see p. 259) it will often limit interpretation of their effect and impair definite conclusions on their relevance for oligochaete occurrence. Substratum Substratum as an ecofactor is a typical example of the complicated interaction of numerous physiographical features which directly or indirectly determine the conditions in a given biotope (Boaden, 1962; Jansson, 1967d, 1968a; Lassèrre, 1967; Giere, 1970, 197le). The size and shape of interstitial space and sediment movement due to agitation by waves and currents can be considered more directly operative. Controlled by these ‘basic’ factors are the sorting of sediment particles, permeability, porosity, and water content. These ‘subfactors’ regulate more indirectly oligochaete distribution, often via chemical and physical intermediaries. Hence, it is only natural that from field studies relatively little evidence can be gained of the direct impact of one particular sediment condition, and the number of conclusive examples is limited. Nevertheless, in many studies the various oligochaete species mentioned are associated with a certain type of substratum. It will be shown in some examples that, in fact, it was not the granulometric composition per se, but a related, less evident feature which was responsible for the worms’ affiliation to a given sediment. In this respect, even the existence of a good correlation to a certain sediment type may be deceptive (Brinkhurst, 1962). When understood as an operational composite integrating the various aspects mentioned above, statements about the substratum as a distributional characteristic have, however, a meaningful ecological significance as shown for some oligochaetes in Table XII. The array of determinants which eventually result in a certain preferred type of bottom varies due to local physiographical and also biotic conditions. This can even lead to apparent contradictions in various publications of the same author reporting different sediment associations of one species (Table XII). Grain size. This directly influences habitat selection and is most relevant in mesopsammic forms for which interstitial space is a crucial condition for life. Most Grania spp. are typical (sublittoral) sand-dwellers, some tropical Phallodrilinae seem to depend on shelly coralline sand, Lumbricillus bülowi (=Fridericia bulbosa sensu Bülow, 1957) and Lumbricillus knoellneri, often to be encountered in the “otoplanid zone” (Knöller, 1935a; Bülow, 1957) are generally restricted to more exposed medium to coarse sands (Pfannkuche, 1977).

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Sediment agitation. For most oligochaetes this acts negatively (Giere, 1977)

TABLE XII Relation of some European marine oligochaetes to sediment types (based on field studies)

Sediment type Mud/silt

Species

Area

Reference

Tubifex costatus

Limfjord, Denmark

Pfannkuche, 1980a

”

Baltic and North Sea shores, Germany

Bülow, 1957

”

Finnish archipelago

Bagge&Ilus, 1973

Peloscolex benedeni

Isle of Man, U.K.

Brinkhurst & Kennedy, 1962

”

Baltic and North Sea shores, Germany

Hagen, 1951

”

North Sea flats, Germany

Linke, 1939

”

Norwegian Sea, shallow subtidal

Erséus, 1976a

”

Firth of Forth, Scotland

McLusky et al., 1980

Paranais litoralis

See above

Pfannkuche, 1980a

”

See above

Hagen, 1951

Clitellio arenarius

Mersey Estuary, England

Fraser, 1932

Fine

Tubifex costatus

Schlei-Förde, W.Baltic

Pfannkuche, 1980c

Fine

Peloscolex benedeni

”

”

Unsorted, rich ” fraction of fine sand

Brittany flats, France

Lassèrre, 1967

Fine

Paranais litoralis

See above

Hagen, 1951

Medium, 160–240 µm

Interstitial Marionina spp.

”

Lassèrre, 1967

Rich in detritus

Lumbricillus lineatus

Schlei-Förde, W.Baltic

Pfannkuche, 1977, 1980c

”

Lumbricillus bülowi

”

”

Sand-mixed

Sand

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”

Paranais litoralis

”

”

Medium to coarse

Tubifex pseudogaster

”

”

”

Clitellio arenarius

”

”

Brackish

Marionina southerni

See above

Hagen, 1951

”

Lumbricillus arenarius

”

”

”

Clitellio arenarius

See above

Brinkhurst & Kennedy, 1962

Stony, gravel

”

Mersey Estuary, England

Fraser, 1932

Gravel

”

See above

Bagge & Ilus, 1973

”

”

”

Hagen, 1951

Coarse sand

”

”

Pfannkuche, 1977

Coarse, 250–2600 µm

”

Brittany flats, France

Lassèrre, 1967

,,

Lumbricillus semifuscus

”

”

”

Lumbricillus lineatus

”

”

”

Lumbricillus viridis

”

”

”

Marionina spicula

”

”

”

Grania macrochaeta

”

”

Species not associated to a particular sediment type Paranais litoralis

Pfannkuche, 1977

Lumbricillus lineatus

Hagen, 1951

Marionina subterranea

”

Monopylephorus irroratus

”

Peloscolex benedeni

Lassèrre, 1967

despite effective adhesive mechanisms developed in many interstitial forms (Giere, 1971c). Giere (1970, 1971e) found Marionina subterranea and M. spicula to avoid the unstable upper levels of a North Sea beach, and observed complete disappearance of formerly rich oligochaete populations (Lumbricillus lineatus, Enchytraeus albidus, Marionina subterranea, Aktedrilus monospermathecus) after a stormy tide with waves heavily agitating the shore sediments. Boaden (1968) concluded from experiments that meiobenthic fauna moved down from surface sediments on Irish beaches when the surge fringe crossed their habitats. For oligochaetes (Marionina spp.) this reaction was corroborated by Meineke & Westheide (1979). Although the absolute numbers of worms may have been under-estimated by them due to the use of the sea water-ice method (p. 178), the trends obtained are interesting. They regularly recorded in a semi-lotic sandy

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slope of the island of Sylt (North Sea) a downward migration at high tide, by which the worms avoided the increased sediment agitation. That the waves’ surge was really evoking the negative reaction could be proved experimentally in columns constantly water-saturated thus simulating permanent high-water conditions but without any wave action. Here, most animals remained in the surface layers (Fig. 12). McLachlan, Erasmus & Fürstenberg (1977) have

Fig. 12.—Fluctuations in vertical distribution of interstitial enchytraeids in a North Sea beach during a tidal cycle and under experimental submersion: LW, low water; HW, high water; Exp. Co., experimental submersed sediment column (modified after Meineke & Westheide, 1979).

also reported cessation from vertical migrations of interstitial fauna in permanently saturated substrata in the absence of agitation (sheltered conditions). Sorting of the sediment. This mainly depends on exposure to waves and Jansson (1967b), Newell (1970), and Hulings & Gray (1976) have stressed its importance as a dominant factor overriding absolute particle size in its relevance for pore space and permeability. According to Sanders (1958), sediments with particles of 180–200 µm diameter are especially exposed to agitation by currents and waves which could explain the avoidance reactions of both macro- and meiobenthic oligochaetes in our own studies on the isle of Sylt (Giere, 1970). 200 µm also seems to be a ‘critical level’ for other mechanical reasons (Wieser, 1959). The high rate at which voids fill up with fine material rapidly excludes life of interstitial forms. On the other hand, larger worms seem to be attracted just by fine sediment, intermediate between silt and ‘open’ sand which would cause mechanical difficulties in burrowing. Removal of the finer fractions by high river flow was regularly connected with reductions in the rich populations of Tubifex costatus and Peloscolex benedeni from the Thames Estuary (Birtwell & Arthur, 1980). (Here nutritive deterioration may also have been involved.)

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In river bottoms, Wagner (1968) found particularly rich tubificid fauna in fine sediments with the fraction <63 µm in diameter being decisive. Locomotor problems were also stressed by Jansson (1967d) who generalized that sliding, ‘nematoid’ movements (in oligochaetes e.g. Marionina subterranea or Grania spp.) highly depend on a certain interstitial space, whereas worms with creeping and burrowing locomotion by means of peristaltic constrictions (e.g. Aktedrilus monospermathecus) only have a slightly developed preference for particularly sorted sands, and thus are more “euporal” (Fauré-Frémiet, 1950). Conrad (1976) showed in experiments that not only the size of interstices, but also the shape of sand grains might determine distribution of interstitial oligochaetes (regrettably no species identification is given) which he found to prefer angular rather than spherical particles. He reasoned that despite the somewhat reduced permeability, diversity of microhabitats and possibly also variety of microbial food sources were enhanced in the geometrically more complex system of angular sand. A similar observation was made by Pfannkuche (1980c) in his Schlei-Förde studies. The number of tubificid species increased with increasing particle size. Relating this pattern to the concomitant increment of microhabitats, he stressed the importance of the “median” sand fraction for establishment of a diverse marine tubificid fauna which basically is confirmed by freshwater species in river sediments (Wachs, 1967). In an interesting approach, Tynen (1972) ranked species number of marine enchytraeids against sediment type and found sand and gravel to harbour the most diverse fauna with Lumbricillus lineatus, L.viridis, Enchytraeus albidus, and various interstitial Marionina spp. to occur most frequently in sand and Lumbricillus reynoldsoni in shingle. The eurytopic nature of L.lineatus is, however, demonstrated by its regular distribution also in shingle. Similarly, Enchytraeus albidus was often encountered in gravel which corroborates Tynen’s (1969) experimental results. Hence, among the macrofaunal enchytraeids, these ubiquitous forms inhabited numerous types of substrata, as do the cosmopolitan Peloscolex benedeni in tubificids and perhaps the widespread interstitial forms Marionina southerni and M.spicula. Water content. The experiments of Meineke & Westheide (1979), mentioned above, have shown water content to be of major importance for oligochaete zonation in shore sediments. During low tide, Marionina spp. preferred the upper, moist, but unsaturated layers of the slope (only 13% water content), a pattern which has already been described for Marionina spp. from Swedish beaches by Jansson (1968a) and from the Schlei-Förde by Pfannkuche (1977, 1980b). Apparently, these low-moisture layers combine sufficient oxygen uptake with good food supply without the risk of desiccation. In subtropical South African beaches, McLachlan (1977a,b, 1978) found a similarly intricate balance between water content, oxygen supply, and desiccation which confined oligochaetes (phallodriline tubificids, Marionina

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Fig. 13.—Field distribution of Aktedrilus monospermathecus in a non-tidal beach in relation to changes in ground-water level (GWL): WL, water line; (redrawn from Giere & Pfannkuche, 1978).

sp.) to the lower, consistently moist zones. In exposed beaches, well drained and with low water retention in the upper layers of the backshore, most worms populated the lower shore (mid- and low-waterline) or, if at highwater level, lived in deeper horizons than they did in sheltered beaches with slow drainage and higher water saturation at ebb tide, where oligochaetes could colonize even the upper layers in the backshore without any risk of extensive water losses. A similarly close adjustment of horizontal and vertical zonation to water content emerges from studies on the interstitial tubificid Aktedrilus monospermathecus (Giere & Pfannkuche, 1978; Pfannkuche, 1980b) from Baltic shores. This species always followed the changing ground-water line of the slope which shifted the population’s position on the slope markedly (Fig. 13, see also Fenchel, Jansson & Thun, 1967; Jansson, 1968a). On the other hand, for many oligochaete species complete water saturation of the sediment acts as a repellant, probably due to the concomitant decrease in oxygen compared to an unsaturated pore system (Schulz, 1955; see also below and p. 252). This is probably the main reason for the marked zonation differences between Enchytraeus albidus and Lumbricillus lineatus. The latter is always closer to the water-line (mainly around high-water level) whereas Enchytraeus albidus, as a semi-terrestrial species and sensitive to oligoxic conditions, seems to avoid permanent submersion (Bülow, 1957) and lives in the backshore, often amongst the reed fringe. This general pattern of the most common European littoral oligochaetes is not invalidated by occasional findings of Lumbricillus lineatus and (juvenile) Enchytraeus albidus below high-water line (Michaelsen, 1927; Seifert, 1938; Pfannkuche, 1980c). These examples from the strand line demonstrate that relation to the water content of the substratum can vary specifically even within the same oligochaete family and, thus, renders generalizations problematical. It can, however, be stated that as a general trend enchytraeids, despite some exceptions (Grania), inhabit the upper slope of sandy shores, frequently exposed to the atmosphere, thus avoiding permanent submersion. Conversely, tubificids and naidids predominantly colonize the fully saturated and less exposed lower reaches and sublittoral substrata, often consisting of fine sediment and muds (Lassèrre, 1967; Jansson, 1968a; Giere, 1970, 1971e; Pfannkuche, 1980b). Turning to permanently submersed or deep-water areas, the relation between oligochaete distribution and sediment structure becomes less camouflaged by interacting salinity-temperature changes and desiccation problems. Cook (1971), in his interesting study on the Cape Cod Bay tubificids, found good agreement between specific grain sizes and tubificid distribution. The oligochaete fauna could be separated into a fine-sand and silt community (8–250 µm) dominated by only two species, Tubificoides intermedius and Limnodrilus medioporus, and a more diverse coarse-sand community (500–2000 µm) consisting mainly of the phallodriline interstitial forms, Tubificoides longipenis and Phallodrilus spp. (Fig. 14). Erséus (1976a) added some Grania spp., Marionina welchi and Lumbricillus codensis to this group. It is the distribution of Peloscolex benedeni which serves as a good example of the

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746

contraditions and difficulties in relating oligochaete occurrence to sediment structure. Usually recorded in rich populations from muddy to fine

Fig. 14.—Distribution of four dominant tubificids, Peloscolex intermedius, Limnodrilus medioporus, Peloscolex benedeni, and Tubificoides longipenis, in Cape Cod Bay, Mass. U.S.A., in relation to water depth and median particle size of substratum (modified after Cook, 1971).

sandy reaches with 80% of sediment <63 µm (Linke, 1939; Hagen, 1951; Bülow, 1957; Muus, 1967; Wharfe, 1975; Birtwell & Arthur, 1980), this species was in Cook’s (1971) work restricted to coarse sand! Also Lassèrre (1967) confirmed from the flats of Brittany the prevailing occurrence of this common form in sandy substrata of a very wide but rather coarse grain size range (300–2000 µm). On the other hand, McLusky, Teare & Phizacklea (1980) and Bagheri & McLusky (in press) found this species restricted to mud and absent in sandy areas. Although it seems that a certain portion of sand is favourable for development of rich P.benedeni populations, there probably is no direct and close relation to a definite sediment type. According to Williams (1972), plasticity in P.benedeni is so large that the “burrower of muddy intertidal beaches” can in gravels

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adopt even a “mesopsammic existence”. Cook (1971) supposed already that restriction of the species to coarse sand might result from a biotic co-determinant such as competitive interaction with Tubifex longipenis. Oxygen supply may also be operative as the real limiting factor in the relatively small and shallow brackish-water bight of the Schlei-Förde, where Pfannkuche (1977, 1980c) related a very differentiated horizontal pattern of oligochaete occurrence to sediment structures (see below) and found Peloscolex benedeni only in the more sandy substrata with a fine fraction (<100µm) not exceeding 20%. Tubifex costatus was mainly found abundant in sediment of grain size >100 µm diameter, while T.pseudogaster mainly occurred in sand of >300 µm grain size (Fig. 15). Discriminatory capacities even within narrower granulometric ranges (between 4 and 250 µm) have been ascribed to some tubificids, studied by Cook (1974) in a shallow bight of Baja California. The fine silts and clay-silts were usually inhabited by T.postcapillatus and Thalassodriloides belli, whereas Limnodriloides barnardi, L.monothecus, and L.verrucosus lived preferably in intermediate silts and fine-sands. Once again, interference by changing availability of organic matter (food) and competitive restriction between

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Fig. 15.—Distribution pattern of dominant oligochaetes in a a-mesohaline bight of the Schlei-Förde (Western Baltic) in relation to sediment type and oxygen conditions (A): for generic abbreviations see text; (from Pfannkuche, 1980c).

Tubifex postcapillatus and Thalassodriloides belli may well account for this sedimentrelated distribution too (see p. 230). The polychaete tubes of Hydroides dianthus (Serpulidae) that were frequently found

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colonized by the tubificids Peloscolex gabriellae and P. intermedius in Delaware Bay (Haines & Maurer, 1980) must also be considered as attractive microhabitats providing not only detritus, silt, and food but also shelter. There is no indication of an interspecific relation between the oligochaetes and the polychaetes themselves. A similar assemblage of a species providing suitable substratum, food, and shelter was Mytilus densely covering the intertidal wave-beaten rocks of the Massachusetts coast and an undescribed Lumbricillus species (Enchytraeidae) which richly populated the voids and crevices underneath the canopy of mussels where detritus and pieces of algal thalli had been trapped between the byssus threads (unpubl. obs.). It becomes evident that the ‘association’ of oligochaete species with certain sediment types (Table XII), is often to be taken as a syndrome of interacting subfactors with ‘sediment’ only as a kind of indicative ‘master-factor’ rather than as a real dependence on grain size per se. On the other hand, an ‘independence of grain size’, stated sometimes for certain species, could be ascribed to some predominant non-sedimentary factors which overshadow an existing impact of grain size. It follows that ascertaining indubitable sediment interactions on oligochaete distribution requires experimental scrutinization under controlled laboratory conditions. Temperature Temperature is often held to be one of the dominant factors regulating distribution of (littoral) oligochaetes (e.g. Lassèrre, 1971b). There are, however, surprisingly few field data on temperature effects although, in tidal flats and on beaches, fluctuating temperatures can be extreme even in boreal regions with an amplitude in emersed sediments several times wider than in the water (Linke, 1939). These fluctuations are restricted to the surface layers and are considerably diminished in deeper strata where both diurnal and annual changes become markedly dampened due to the ‘buffering capacity’, directly depending on drainage and evident also from the steep vertical salinity gradients already mentioned (Renaud-Debyser, 1963; Jansson, 1966a, 1968a,b; Giere, 1970, 197le). In wrack beds, another biotope often harbouring abundant oligochaetes, the dampening effect is even more pronounced (Backlund, 1945; Schulz, 1955). Even in the sediment layers close to the surface, temperature recordings above 25 °C are rare in temperate climatic zones. In subtropical and tropical flats, however, the uppermost sediment may often reach 35 °C and more during low tide (Jansson, 1966b; Giere, 1977). On the other hand, not only in polar regions, but also in areas like the North Sea and Baltic shores, temperatures below zero can occur for longer periods and often the upper layers are frozen solid (Linke, 1938; Jansson, 1967c, 1968a). Penetration of frost is deepest in sandy beaches (Jansson, 1968b) and least in wrack (Backlund, 1945; pers. obs.). Reports on a possibly adverse impact of these extreme temperatures on the field distribution indicate a dependence on ambient conditions. In general, most littoral oligochaetes can apparently endure temperatures below zero and even survive for some time in the frozen sediment. This refers both to enchytraeids and tubificids, to macro- and meiobenthic species (Stephenson, 1930). Lumbricillus aegialites moves on snow on top of seaweed (Stephenson, 1925); Peloscolex benedeni recovered completely after being

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frozen solid for several tides in a mud flat (Linke, 1939). Marionina southerni was found alive in frozen sand (Jansson, 1968a). Living oligochaetes were encountered by Leloup & Miller (1940) at −2°C in a brackish lagoon. The ubiquitous Lumbricillus lineatus and Enchytraeus albidus populating wrack beds in every season (Backlund, 1945) remained undamaged after occasional freezing of sand plus seaweeds below −10°C (Schulz, 1955) and could be thawed out in good condition from a batch of Fucus thalli accidentally deep-frozen beyond −20°C for at least one day (Hauschildt, pers. comm.). The considerable ability of most oligochaetes to endure sudden low temperatures and reduced salinities is evidenced also by attempts to extract them by the sea water icemethod; many individuals stay coiled up and ‘stiff’, but alive in the sample core which results in completely erroneous quantitative data (pers. obs.). Worms, whose natural habitat is snow and ice, like Mesenchytraeus solifugus and M.gelidus (see Stephenson, 1930), are not known from brackish or marine (polar) environments. It has already been pointed out by Michaelsen (1927), however, that this good survival is not unaffected by environmental conditions. He explained the short survival of Enchytraeus albidus in frozen water, compared with high viability in frozen wrack, to the lack of sufficient oxygen under the former condition. This may be decisive also for enchytraeid survival in sheltered shores with anoxic layers close to the surface (Giere, 1970, 1971e). Here, oligochaetes which tend to avoid temperature extremes by downward migrations into the less rigorous subsurface horizons (see p. 264), are soon stopped by insufficient oxygen in combination with developing hydrogen sulphide (see below) and, thus, are often bound to persevere in the top centimetre layer, exposed to long-lasting frost, often combined with the concomitant problem of desiccation. These adverse conditions, prevailing along the North Sea shore in the harsh winter of 1978/79, caused a drastic reduction in abundance of enchytraeid populations and retarded the growth of the worms which were found sluggish, unusually small, and in a generally bad condition. Corresponding conditions in extreme warmth are even more stringent; at high temperatures, survival of marine oligochaetes is less (see p. 263), and oxygen demand is considerably enhanced. Even in temperate or boreal regions, summer conditions can cause an ecological situation in surface layers harmful for enchytraeids despite other favourable ecofactors. In these regions, although not yet confirmed, the common paucity of enchytraeids in surface layers of sandy flats may be explained by temperature if the possibility of withdrawal into deeper levels is hampered (Giere, 1970). In a subtropical Bermudian flat, the adverse impact of high temperatures (35°C) on interstitial oligochaetes has been demonstrated (Giere, 1977; Giere & Pfannkuche, 1978), see p. 264). The oxidized surface layer only 1 cm thick in the mid-tide areas of a flat did not provide sufficient protection from extremes of temperature (and salinity) to enable a persistent colonization by oligochaetes. As soon as the increased oxygen supply even in deeper horizons was granted (high tide areas), the worms could obtain protection from adverse temperatures by downward migration and, thus, richly populated these regions. Here, an adverse ecological background can evoke (high) temperature as the decisive factor for oligochaete occurrence. Turning now to the influence of temperature on long-term population developments, we have to focus attention on seasonal temperature changes. Although low temperatures have not been shown to exclude oligochaete life, winter conditions are generally held

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responsible for the seasonal population breakdown in Naididae and also for the considerable reduction in enchytraeid and (interstitial) tubificid populations (Nielsen, 1955b; Pfannkuche, 1979; 1980b,c). Besides the enhanced mortality, winter temperatures especially seem to cause in many enchytraeid species a general retardation of growth and a cessation of maturation processes (Reynoldson, 1943, 1947a; see p. 195 ff and Table III) resulting in smaller persevering stocks of submature and juvenile worms, as was drastically demonstrated after the harsh winter 1978–1979 along the North Sea shores. Enchytraeids in wrack beds or sand underneath the seaweed fringe do not follow this general pattern to the same degree because of the high insulating capacity and warmth retention of plant matter (Schulz, 1955). Hence, mature E.albidus, Lumbricillus lineatus, and also Marionina spicula could be encountered here even in winter (pers. obs.; Backlund, 1945). Reynoldson (1947a) even reports from sewage beds maximal abundance and high breeding activity of Enchytraeus albidus at winter temperatures. During spring and summer, population development by increased reproduction is clearly stimulated (Hagen, 1951; Giere & Hauschildt, 1979). Increase of temperature was found to trigger the onset of breeding in Tubifex costatus which could contribute to the positive correlation between biomass and temperature in this common species (Birtwell & Arthur, 1980). On North Sea and Baltic shores, maximal abundances were regularly recorded after periods of warm weather (Giere, 1970, 1971e, 1975) provided disturbance of sediments remained slight and the water content high enough. This parallels the statement of Nielsen (1955b) who generalized that terrestrial enchytraeids have summer maxima in regions with oceanic climate. Whether the spring maximum and summer decline, common in naidids (see Fig. 5, p. 194; Heip, 1971), is mainly triggered by temperature or depends on availability of food, is as yet unclear. Considering the eurythermic habit of many oligochaete species and the ever-increasing number of species from tropical seas (Jamieson, 1977; Giere, 1979b; Erséus, 1980a), the statement of Timm (1980) that reproduction in tubificids is stenotherm, adapted to cool temperatures is perhaps not justified. Salinity Even more frequently than sediment structure, salinity has been underlined as the environmental factor governing the distribution of marine oligochaetes. Especially in brackish-water intertidal areas and estuaries, where it fluctuates most due to tides, storms, evaporation, and ground water influx, salinity seems to control the distribution of these worms (Brinkhurst & Kennedy, 1962; Laakso, 1969; Moroz, 1977; Birtwell & Arthur, 1980; see p. 253 ff). Comparing the numerous salinity range data and diagrams of littoral oligochaetes from different regions (Michaelsen, 1927; Knöllner, 1935a; Hagen, 1951; Bülow, 1957; Moroz, 1974; Pfannkuche, 1974, 1977, 1980a,c), considerable variations become evident for the same species (Table XIII), probably based on the local interdependence of salinity with other factors

TABLE XIII

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Selection of field data on the salinity range of some brackish-water and marine oligochaetes

Species Tubifex costatus

Salinity range ( ‰S)

Area

1–19 Thames Estuary

Birtwell & Arthur, 1980

4–30 German shores

Hagen, 1951

2–16 Schlei-Förde

Pfannkuche, 1974

1–23 Kiel Bight

Knöllner, 1935a

4–7 Finnish archipelago 2–36 German shores 20–28 French Atlantic shores, English Channel Clitellio arenarius

9–34 ,, >6 Finnish archipelago >5.7 See above

Nais elinguis

Paranais litoralis

Reference

Bagge & Ilus, 1973 Michaelsen, 1927 Lassèrre, 1967 ,, Laakso, 1969 Bagge & Ilus, 1973

11–21 ,,

Knöllner, 1935a

14–30 ”

Hagen, 1951

6–34 ,,

Michaelsen, 1927

0–30 ”

Hagen, 1951

2–16 Schlei-Förde

Pfannkuche, 1980c

0–24 See above

Knöllner, 1935a

6–13 Zuider Sea

De Vos, 1922

0–30 See above

Hagen, 1951

8–14 ,,

De Vos, 1922

2–16 ,,

Pfannkuche, 1974

>1 ,, 0–36 ,,

Laakso, 1969 Michaelsen, 1927

15–33 ,,

Lassèrre, 1967

24–29 Delaware Estuary

Watling, 1975

such as temperature, sediment structure, emersion, and also pollution (see p. 281). In such fluctuating conditions, irregular changes of salinity due to extreme evaporative water loss

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or sudden rainfall seem to determine directly oligochaete distribution in the intertidal. In a temperate climate, the dominant shore species are, however, well adapted due to their wide range of short-term tolerance (p. 272). Moreover, older reports on extreme surface salinities (e.g. Gerlach, 1954) did not consider the steep vertical gradient which enabled the animals to live in near-surface layers in fairly ‘normal’ conditions. Probably, this micro-stratification, indicating a substantial dampening effect of the substratum, could not be adequately detected by the older measuring devices. Many recent studies (Bush, 1966; McLachlan et al., 1977; pers. obs.) underline this ‘buffering’ capacity of littoral sediments, high enough to maintain viable conditions even in extreme salinity situations. Only in subtropical and tropical flats, could direct limitation by salinities, often exceeding 40‰ S, become an important distributional factor (see p. 263), as shown for Aktedrilus monospermathecus from a Bermudian cove (Giere, 1977; Giere & Pfannkuche, 1978). The high variability of salinity ranges originating from field studies limits the value of simply listing all the data available for the various species. Therefore, it is attempted here to point out some general traits of salinity-induced effects on the distribution of oligochaetes. Comparative studies in estuarine areas and relatively stable brackish waters (Diaz, 1980) suggest that it is rather the unstable conditions within the salinity regime which indirectly control oligochaete distribution. Diaz compared the occurrence of brackish-water tubificids in a poikilohaline American tidal estuary with that in a homoiohaline Finnish archipelago and found in the unstable and polluted estuarine environment a reduced species diversity but high population numbers. He compared this with results from Bagge & Ilus (1973) who had reported a richer tubificid diversity with low abundance from the Baltic waters. Pfannkuche (1980c), based on his own studies and investigations of Laakso (1969), Arlt (1973, 1975), Ankar & Elmgren (1975, 1976), and Leppäkoski (1975), ranks oligochaetes as important members in β-mesohaline areas, but reports maximal values from the a-mesohalinicum (see also Dahl, 1960; Muus, 1967). He suggests, however, many quantitative results for oligochaetes given in literature may be questionable due to inadequate sieving, non-consideration of life cycles, and seasonal population fluctuations. Although the value of these comparative results is limited because the biotopic conditions and the methods applied differ considerably, another aspect of oligochaete occurrence in relation to salinity, namely their distribution in brackish-water gradients similarly underlines the relevance of salinity fluctuations. Comparing field data on salinity limitation of oligochaetes from estuaries (e.g. Schlei, Elbe: Pfannkuche, 1974; Pfannkuche, Jelinek & Hartwig, 1975; Thames: Birtwell & Arthur, 1980) with values from the less fluctuating Baltic (Knöllner, 1935a; Hagen, 1951; Laakso, 1969) and Black Sea (Dnieper Bug-lagoon: Moroz, 1974), there often emerges a more restricted oligochaete occurrence in tidal estuaries relative to the more stable brackish seas and lagoons. Hence, the distributional limits of single species in salinity ecoclines largely depend on local environmental conditions as an integrated complex (see for instance interaction of salinity, temperature, and oxygen with occurrence of Tubifex costatus in the Thames Estuary, Birtwell & Arthur, 1980) and are, per se, less meaningful to report than occurrence of ‘species groups’ which relate to salinity categories. Based mainly on Michaelsen’s (1927) system, a euryhaline freshwater group is usually separated from the euryhaline marine forms by genuine brackish-water species which rarely populate

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limnetic or euhaline regions. The opposite end of this gradient is occupied by relatively stenohaline marine species rarely entering brackish water (Hagen, 1951; Bülow, 1957; Pfannkuche, 1974, 1980c). Knöllner (1935a) diversifies this system by introducing subdivisions without altering its principal categories. A recent scheme of distributional classes, originally referring particularly to estuarine gradients (Boesch, 1977), has been modified by the authors specifically for the distribution of some characteristic European temperate-water oligochaetes in salinity gradients, subdivided into species of the upper eulittoral, mostly emersed shore (Fig. 16), and normally littoral sediments (Fig. 17). Only the stenohaline marine group includes many species which

Fig. 16.—Associations of common boreal shore oligochaetes distributed along a hypothetical salinity gradient (classification modified after Boesch, 1977): -----irregular occurrence; I and II in L.rivalis indicate possible sibling species.

leave the littoral fringe and regularly occur in the open oceanic benthos. Separation of typical opportunists from the bulk of euryhaline marine forms accounts for the fact that especially in ecological stress situations, the oligochaete fauna is often widely dominated by few successful and adaptable species often displacing other forms and occurring in extremely wide salinity ranges. It is interesting to note that among the brackish-water group small interstitial species have a particularly large share and occasionally enter even

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freshwater sediments by invasion via the ground-water system (“Küstengrundwasser”, see Knöllner, 1935a,b). Among the limnetic forms living submersed (Fig. 17) enchytraeids are usually absent while naidids and tubificids dominate. Enchytraeus albidus, abundant worldwide from garden soil and freshwater to hypersaline sediments, belongs to the rare animals apparently not restricted by any salinity (see tolerance experiments, Table XV). Nevertheless, the cytological background in this species should be scrutinized in order to ascertain genetic unity of the various populations. Corresponding studies with Lumbricillus rivalis revealed a cluster of several morphologically identical sibling species (Christensen & Jelnes, 1976, see p. 273), which accounts for the undisrupted occurrence of this ‘species’ in fresh water, brackish water, and sea water (see similar results on opportunistic polychaete ‘species’ like Capitella capitata and Ophryotrocha spp. (Grassle, 1980). The threshold between limnetic and marine populations is reported by the authors to lie in the oligohalinicum. Examination of Figures 16 and 17 shows surprisingly good agreement

Fig. 17.—Associations of common eulittoral and sublittoral boreal oligochaetes distributed along a hypothetical salinity gradient (classification

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modified after Boesch, 1977) : -----irregular occurrence.

between the classical biological salinity-thresholds of the “Venice System” (1958) and the limitations of oligochaete occurrence; the 25‰-, 18‰-, and 5‰-borders are especially well expressed. Brackish-water endemists are bound to the mesohaline salinity range. Limnetic forms (most naidids) do not proceed beyond the oligohalinicum into higher saline water. Stenohaline marine species are limited to the fully marine and upper polyhaline zone. Marine euryhaline oligochaetes, although living also in euhaline waters, seem to thrive best in the lower poly- and mesohalinicum. Hence, the thresholds, resulting originally from general benthos and also planktonic studies (Remane, 1940) agree well with the occurrence of littoral oligochaetes. Regarding their extreme ecological flexibility, it is not surprising to find the opportunistic species not fitting into this salinity classification pattern but residing in almost the whole salinity cline. It is, besides other features, this widespread occurrence which justifies their ranking as euryhaline opportunists in contrast to the euryhaline marine species. Alkalinity In studies on the marine biota, pH is usually considered a conservative environmental factor, well buffered by various chemical processes and of insignificant effect for benthic animals (Ben-Yaakov, 1973; Wieser, Ott, Schiemer & Gnaiger, 1974). In temperate regions, this general picture holds true even in tidal flats exposed to light and, thus, high assimilation rates of epibenthic algae. Here, pH values above 8.5 have rarely been recorded in the surface layers of sediments which is only slightly higher than alkaline conditions in the ambient sea water. Hence, no adverse effect on oligochaetes inhabiting these layers is to be expected and none could be demonstrated in tolerance experiments. Studies from a Bermudian flat showed, however, that in a warmer climate the dense diatom populations of sheltered calcareous surface sediments can occasionally give rise to such an intensive assimilation activity that pH values easily exceed the 9.0 threshold and can be as high as 9.6 (Wieser et al., 1974; Gnaiger, Gluth & Wieser, 1978). Tolerance experiments (p. 264) verified that this extreme alkalinity caused a deterioration of viability of the interstitial tubificid Aktedrilus monospermathecus especially if combined with additional stress factors like salinity around 40‰ and temperatures of 35° C. These conditions occur regularly in subtropical and tropical shores on hot sunny days, and certainly contribute to the reduced population density of these oligochaetes in habitats where anoxic subsurface layers prevented avoidance by vertical migrations (Giere, 1977: Fig. 1). Hence, in (calcareous) subtropical and tropical shore sediments, extraordinary rich in marine oligochaetes (Jamieson, 1977; Erséus, 1980a), this extreme increase, apparently, is not an unusual phenomenon. Here, alkalinity has to be taken into account as a factor contributing effectively to the distribution pattern of oligochaetes. Light Light, a factor directly and indirectly (via plant assimilation) highly decisive for the occurrence of marine fauna (Segal, 1970), has been almost completely neglected in

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studies of the distribution of marine oligochaetes, probably due to the often observed and generally recognized concept that these endobenthic worms are negatively phototactic. This negative response was confirmed by Welch (1914) for Lumbricillus lineatus. Hagen (1951) stated a generally developed negative light sensitivity in littoral oligochaetes. This is also commonly found in most studies on vertical distribution of shore oligochaetes where high insolation may contribute, as well as other factors, to the usual paucity of oligochaetes in surface layers. Both in tubificids and enchytraeids, the prostomium and pygidium are the regions most susceptible to light. The study of Studentowicz (1936) on Enchytraeus albidus proved experimentally that negative phototaxis could outweigh attraction by food and also thigmotactic responses. Only oxygen deficiency and hydrogen sulphide (see below) seem to override retractions from light. In hypoxic substrata the worms accumulated at the surface even in bright light. Studentowicz also showed that the species can discriminate between slight differences in light intensities. According to the authors’ personal ecological field and laboratory experience one can probably assume that all other marine enchytraeids and tubificids behave in a similar way. For the naidid worms at least, mostly equipped with eyes and often found to swim in the epibenthic water layers, the general assumption that light repels the animals simply cannot, however, be accepted. Regarding the abundance and nutritive importance of the brackish-water naidids (see p. 193 and p. 213), detailed studies on their response to light as a cyclic factor regulating diatom blooms and possibly interfering with activity rhythms in predative fish are urgently needed. Considering the studies on many polychaetes, where light has been found to influence maturation processes and to synchronize reproductive activities, it would not appear to be meaningless to initiate pertinent studies also for oligochaetes. Whereas Lassèrre, (1967, 1970) for Marionina spp. could not confirm clear daily vertical migrations, Faubel (pers. comm.) has obtained results on sublittoral populations of Grania postclitellochaeta and juvenile Lumbricillus lineatus (?) from coarse sands in Kiel Bight (western Baltic Sea, 8– 20 m depth) which seem to indicate a vertical migration pattern. During light periods, rich populations of the worms were encountered in the uppermost 3 cm of the sediment, whereas these layers were regularly found abandoned at night. Oligochaetes developing dark pigment layers in their epidermis as a shelter against detrimental isolation like Mesenchytraeus solifugus and M.gelidus from glacier areas (Stephenson, 1930; Tynen, 1970) have not been reported from marine biotopes. Oxygen and hydrogen sulphide The decisive rôle, oxygen and (conversely) hydrogen sulphide play in the distribution of marine oligochaetes has been mentioned earlier for wrack-bed forms and more clearly demonstrated for mesopsammic than for mud-dwelling species. Enchytraeus albidus lives only in aerobically decomposing wrack and dies in anoxic conditions (Backlund, 1945). It aggregates around air enclosures in moistened soil and, in hypoxic conditions, moves up to the surface layers tolerating even unfavourable impact of bright light (Studentowicz, 1936). These field observations in relatively natural habitats are in some contrast to experiments from which a pronounced ability of E. albidus to survive in anaerobic conditions for many hours can be inferred (Ivleva, 1960). This is in contrast to

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earlier results of Harnisch (1942) who determined in cell homogenates of this species a very high oxygen consumption. According to Welch (1914) Lumbricillus lineatus has an “abundant” oxygen demand, although Knöllner (1935a) and Hagen (1951) reported rich populations in brackish ponds strongly smelling of hydrogen sulphide (see below). For psammobiotic oligochaetes, in general, and particularly for enchytraeids, they concluded, as has been confirmed by later studies (McLachlan, 1977a,b, 1978; see below), that the distribution was closely related to the oxygen regime (Stephenson, 1930; Giere, 1970, 1971e). In experiments, Nielsen (1961) underlined the high oxygen demands of terrestrial enchytraeids. More specifically Jansson (1967b, 1968a) ranked oxygen, recorded as “availability”, of prime importance for the restriction of interstitial oligochaetes (Marionina spp., Aktedrilus monospermathecus) to the oxidized surface layers. Similarly, Giere (1970, 1971e) found close parallels between varying oxygen content (measured as “availability” and redox potential) in heterogeneous beach sediments and vertical and horizontal distribution patterns of oligochaetes. The important rôle of oxygen can be most clearly shown in sheltered beaches with a well-developed redox potential discontinuity (RPD)-layer underneath oxidized surface layers. A comparative study between this type of shore from the German coast and well-oxygenated, exposed Scottish beaches (Giere, 1973) revealed a vertical limitation of both macro- and meiobenthic oligochaetes by oxygen diffusion rates of 1×10−7g·cm−2·min−1or a redox potential of 0 to +50mV, usually recorded in the vicinity of the visible bordering zone between light brownish or yellowish to grey or black sediments. If in sheltered sediments with rich organic contents these oligoxic layers are shifted closer to the surface by anaerobic decomposing processes leaving, for example, only a 1-cm thin surface layer of sand oxygenated, the abundance of enchytraeids and psammobiotic tubificids will be greatly reduced. In the few mm’s of oxidized sand in a Bermudian sand flat (Giere, 1977) almost complete absence of A. monospermathecus and Marionina spp. (populating in rich numbers the deeper oxidized highwater line) was caused mainly by anoxia in subsurface sediments. In brackish habitats, these thin oxidized horizons harbour often large numbers of naidids living predominantly in the sediment-water interface. The distribution of species like Nais elinguis, Paranais sp., and Chaetogaster sp. was also found restricted to oxygenated sediment layers (Fenchel & Jansson, 1966). In tubificids, the distributional impact of oxygen is better investigated, but also somewhat confusing. As in limnetic environments, in many marine studies tubificids like Peloscolex benedeni and Tubifex costatus are recorded from sediments with extremely low oxygen tensions (Knöllner, 1935a; Muus, 1967; Hunter & Arthur, 1978; Birtwell & Arthur, 1980). Hence, in sediments with reduced lower layers these species are usually positioned at a greater depth than enchytraeids (Giere, 1970, 1971e). The detailed reaction of the large, common marine tubificids T.costatus and Peloscolex benedeni to oxygen deficiency seems, however, to differ depending on specifically varying oxygen demands. Hunter & Arthur (1978) and Birtwell & Arthur (1980) reported from the Thames Estuary that Tubifex costatus followed the seasonal changes in oxygenation of sediment layers whereas Peloscolex benedeni, “able to tolerate anaerobic conditions”, remained in its layer even when the black oligoxic or anoxic horizons moved upwards. This resulted in a greater general penetration of this species into the black layers of the bottom. This distributional difference was in good agreement with tolerance experiments

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(Birtwell & Arthur, 1980; see p. 265) where P.benedeni survived for longer periods in anoxic conditions than did Tubifex costatus. The distribution of T.costatus in Danish brackish-water areas (Dahl, 1960) confirms the need of this species for a certain amount of oxygen, as it occurred only sparsely where oxygen level was low, whereas Peloscolex benedeni was found in polluted bottoms “stinking of hydrogen sulphide” quietly buried in the substratum with its tail sticking out. Dahl (1960) supposes this species to have a welldeveloped intestinal respiration like other (limnetic) Tubificidae (see Alsterberg, 1922). For Tubifex costatus, Hunter & Arthur (1978) found a larger susceptiblity for oxygen deficiency in juveniles than in adults which could explain the greater incidence of newlyhatched worms in the superficial, better oxygenated sediments. Van Hoven (1975) found in experiments with naidids and tubificids a well-developed preference response to richly oxygenated water. He concluded that particularly his naidid species, Branchiura sowerbyi, Dero nivea, requiring a well oxygenated environment, will meet strong competition and predation, while Ilyodrilus templetoni and Limnodrilus hoffmeisteri will hardly be limited in their natural habitats by low oxygen tensions and, so, in polluted waters with their reduced competition represent “indicator species”. Although based only on one location, a similar rank is given by Subba Rao & Venkateswara Rao (1980) to the Indian tubificids Monopylephorus indicus and M.waltairensis which dominated in “black mud” of an Indian harbour “emanating hydrogen sulphide”. In a detailed study on tubificid occurrence, Pfannkuche (1980c) showed that for Tubifex costatus, T.pseudogaster, and even for Peloscolex benedeni the vertical distribution so closely followed the oxygen content in the sediment column that he could correlate the depth of oxygenated layers (redox-potential+200 mV) with maximal depth of occurrence of tubificids (Fig. 18). This referred also to the seasonal changes in depth of the

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Fig. 18.—Seasonal changes in sediment depth distribution of Tubificidae in relation to position of oxygenated sediment layer (indicated by the +200 mV-threshold) in the Schlei-Förde (Western Baltic) (modified after Pfannkuche, 1980c).

oxygenated layer, indicated by fluctuations of the RPD-horizon. Pearson & Stanley (1979) concluded from an intensive sampling programme in Scottish lochs that in sediment with high input of organic matter and redox potentials between –150 mV and 0 mV (at 4 cm depth), indicating predominantly anoxic conditions, the fauna was dominated by “small annelids” and P. benedeni in particular. At values around 0 mV and higher, the “larger annelids” (polychaetes) took over but their total abundance decreased. Summarizing, it can be generalized that for all interstitial forms and for the bulk of Enchytraeidae, in particular, oxygen is a distributional master factor (probably only comparable with food, see p. 247) which predominantly restricts these groups to well drained layers of sandy substrata (McLachlan, 1977a,b, 1978). Generally, the Tubificidae, living more frequently in sheltered and submersed muds or fine sands, are somewhat less oxygen-demanding. Few specialists even inhabit occasionally anoxic sediments. This distributional pattern emerges also from their vertical zonation, relating depths of occurrence with redox-potential. In the emersed hygropsammal, enchytraeids and interstitial tubificids regularly populate only the upper, highly oxygenated layers with

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Eh-values around+200 mV and more (see Figs 22 and 24; Giere, 1973), whereas the larger tubificids in cores from submersed habitats often occur in horizons well below the surface in a low-oxygen regime of 0 to +100 mV (Fig. 19; Pfannkuche, 1980c).

Fig. 19.—Vertical distribution of some tubificids related to the Eh-regime in the sediment of the Schlei-Förde (western Baltic): for generic abbreviations see text; (modified after Pfannkuche, 1980c).

Other distributional effects than oxygen may, however, influence this pattern. The sediments around the RPD-layer have been shown to harbour especially rich bacteria populations (Yingst & Rhoads, 1980), possibly an attractive factor which may result in aggregations of oligochaetes able to tolerate, at least intermittently, very low oxygen tensions like many tubificids (see p. 265). For the common limnetic tubificid Limnodrilus

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hoffmeisteri, Fisher & Beeton (1975) pointed out that depth distribution, although largely depending on oxygen supply, could be modified by nutrition; in order to take advantage of rich food sources, the worms may accept even sub-optimal oxygen conditions. The reduced oxygen consumption in a mixed assemblage of three tubificid species compared with monospecific cultures (Brinkhurst, Chua & Kaushik, 1972) represents a highly interesting metabolic interaction of associated oligochaetes. That in marine species periods of life are endured in corresponding situations is indicated by the fact that Tubifex costatus living in sediments with low oxygen supply has a much lower weight than corresponding ‘normal’ populations (Birtwell & Arthur, 1980). This compromise regarding oxygen conditions could explain some contradictions between oxygen-dependent general distribution and occasional local aggregations of marine oligochaetes in the field or results from tolerance experiments (see p. 265). Birtwell & Arthur (1980) found T. costatus in the Thames Estuary, although “relatively intolerant of anaerobic conditions”, also in places where often extremely low oxygen tensions prevailed. They suggested, however, that oxygen uptake by the worms is periodically changing and facilitated in periods of tidal emersion. This demonstrates again that despite the substantial rôle oxygen plays as a major distributional factor, oligochaete distribution in nature is the result of the interaction of many factors with time as an additional dimension. The repellant effect of anoxic layers on distribution of marine oligochaetes is strongly enhanced if oxygen deficiency is combined with development of hydrogen sulphide. Interstitial forms in particular avoid even traces of hydrogen sulphide (Lassèrre & Renaud-Mornant, 1973), whereas especially the large tubificids Peloscolex benedeni, Monopylephorus indicus and M. waltairensis, as well as the enchytraeid Lumbricillus lineatus have been recorded from sheltered and polluted localities smelling of hydrogen sulphide (Knöllner, 1935a; Hagen, 1951; Dahl, 1960; Muus, 1967; Subba Rao & Venkateswara Rao, 1980). Tolerance of hydrogen sulphide has not yet been experimentally ascertained for marine oligochaetes, as it has for other invertebrates (Theede, 1973a; Caldwell, 1975). Lassèrre & Renaud-Mornant (1973) report, however, from experiments that 7mgH2S·1−1 is fatal for Marionina achaeta. Detailed distributional analyses revealed that at least deeper horizons persistently enriched with hydrogen sulphide were always devoid of any marine oligochaete (Hagen, 1951; Giere, 1970, 1971e, 1973, 1977; Pfannkuche, 1980c). This is in some contrast to limnetic sediments where species like Limnodrilus hoffmeisteri and Tubifex tubifex are relatively independent of the oxygen and hydrogen sulphide regime (Alsterberg, 1922; Aston, 1966; Walker, 1970; Pfannkuche, 1980c). This capacity may refer to the fact that tubedwelling, common in freshwater tubificids, is not known from the marine species, since they are free-burrowing (Virnstein, 1979; see p. 245). Hence, their bodies would be directly exposed to the poisonous impact of hydrogen sulphide. In the H2S-layers of muds, often separated from the atmosphere or from oxygenated water by only a few millimetres of oxygenated substratum or algal mats, where specialists have been found, the situation may be somewhat different. The oligochaetes regularly found here (Paranais litoralis, Lumbricillus lineatus, Peloscolex benedeni, Knöllner, 1935a; Bülow, 1957) could easily utilize the oxygen from the surface and, hence, live in an actual microhabitat not persistently and completely anoxic. Although possibly living close to

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their tolerance limits for hydrogen sulphide, as has been tested for other organisms from comparable habitats (Caldwell, 1975), they would have little competition in habitats with optimal nutrient supply. This may account for the somewhat confusing findings of these species in sulphide-smelling bottoms. Whether in these oligochaetes a ‘sulphide detoxificating system’ exists, corresponding to the conditions in some truly thio-biotic interstitial forms (Powell, Crenshaw & Rieger, 1979), is still unknown. In any case, a thio-biotic marine oligochaete living obligatory and permanently in anoxic sediments is hardly to be expected. BIOTIC FACTORS Bioturbation and trophic aspects In addition to the major abiotic factors which, as described above, influence the distribution of marine oligochaetes, biotic effects have to be considered. Among these, bioturbation of sediment by the worms is in a connecting position between biotic activities and edaphic, physical, and chemical consequences. In freshwater habitats, the effect of the oligochaete’s burrowing activity on stratigraphy, transformation of chemical compounds and mineralization of pollutants is fairly well documented (Ravera, 1955; Appleby & Brinkhurst, 1970; Davis, 1974a,b; Davis, Thurlow & Brewster, 1975; further literature therein; McCall & Fisher, 1980). In coastal and marine areas, the relevance of bioturbation has only fairly recently been recognized (Cullen, 1973; Rhoads, 1974; Driscoll, 1975), and hardly anything is known specifically from areas populated with marine or brackish oligochaetes. Despite often similar population sizes, direct comparison with limnetic oligochaete assemblages is inadmissible as the marine species are not tube-dwellers feeding in certain preferred strata (2–7 cm below the surface: Brinkhurst & Kennedy, 1965; 3–6 cm: McCall & Fisher, 1980; 2–8 cm: Pfannkuche, 1981) and regularly depositing their faeces on the sediment surface, but they move freely in the substratum at varying depths as vagile burrowers or endo-/mesopsammic forms, randomly distributing their faeces. For these biological reasons, and also because of their smaller size and the enhanced competition with polychaetes, we probably have to assume an attenuated effect of oligochaete bioturbation in the marine field (Diaz, 1980). Nevertheless, based on general trends reported from limnetic studies, and on some results from pertinent estuarine oligochaete investigations one may assume that pelletization of ingested sediment and re-working by the worms’ creeping activity leads to a compacting of the sediment, mechanical stabilization, and probably increase in microhabitats. The inherent mellowing of the sediment enhances permeability, influx of oxygenated water and, thus, depth of RPD-layer (Davis, 1947b; Wood, 1975; McCall & Fisher, 1980). The close correlation between the depth of oxygenated sediment layer and population size described by Birtwell & Arthur (1980) for brackish habitats of the Thames Estuary (Tubifex costatus), has been expressed as a regression line by Schumacher (1963) from limnetic flats of the Elbe Estuary (probably Tubifex tubifex, Limnodrilus spp.), and by Pfannkuche (1980c) for three tubificid species in the brackish Schlei-Förde (see Fig. 20). The feeding

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Fig. 20.—Linear regression between maximal depth of tubificid recordings and extension of oxygenated sediment layer (indicated by the+200 mVthreshold) in two stations of the a-mesohaline part of the SchleiFörde (Western Baltic) (from Pfannkuche, 1980c).

activity of the worms will support digging of detrital material from the surface into deeper layers and subsequent bacterial decomposition. Enhanced bacterial activity in combination with good oxygen conditions, however, will attract bacterivorous oligochaetes and favour population growth (Yingst & Rhoads, 1980; see p. 208). It results in a self-created magnification of a sediment layer biogenically meliorated for oligochaete settlement via physico-chemical and nutritive processes (Gerlach, 1978; Reise, 1981). Additional biogenic structures like polychaete or crustacean tubes and plant roots would further enhance spatial partitioning, sediment oxygenation, and microhabitat diversity and could increase oligochaete settlement, especially of meiobenthic forms (Bell, Watzin & Coull, 1978; Reise & Ax, 1979; Alongi & Tietjen, 1980). Estimates of the amount to which brackish and marine oligochaetes control this feedback relationship (Driscoll, 1975) and chemical cycling, and cause intensive circulation of the upper sediment layers, as stated for many freshwater sites (Wood, 1975; McCall & Fisher, 1980), must still be tentative. Considering the often vast oligochaete numbers even in

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shallow, eutrophicated coastal areas (see p. 199) the beneficial effect of oligochaete populations should not be under-estimated even for marine environments (Aller, 1980). Assuming 30 ml of sediment reworked per year per worm, as given by Diaz (1980) for an American tidal estuary, and an average population size of 4000 Tubificoides gabriellae·m−2 (Virnstein, 1979), there would result 120 1 of displaced sediment·m−2·yr−1. This would mean that the entire upper 5 cm, the preferred settling depth of T. gabriellae (Diaz, 1980) would be reworked more than twice a year by the worms’ activity (in Toronto Harbor: 4–12 times according to Appleby & Brinkhurst, 1970). Birtwell & Arthur (1980) reported for the Thames Estuary 7 kg·m−2 displaced per day mainly by Tubifex costatus which corresponds to 2·5 tons sediment·m−2·yr−1. Even if these data are somewhat speculative, they illustrate that also in the marine environment biogenic circulation of the substratum and self-creation of favourable habitat conditions initiated through oligochaetes should be taken into account. Trophic effects Trophic effects on occurrence of oligochaetes, mentioned already as relevant in bioturbation processes (digging of deposited detritus, increase in microbial activity) have to be discussed as another biotic factor determining the distributional pattern. Beyond the common but rather vague knowledge that oligochaetes occur mainly in layers rich in organic substances, few detailed studies based mainly on field observations relate the aggregative occurrence of marine oligochaetes to the ambient sediment and its food content. Decomposing seaweeds certainly are the distributional basis for the rich stocks of enchytraeids (mainly Enchytraeus albidus, Lumbricillus lineatus) aggregated by the abundant food supply in the wrack beds (Welch, 1917; Moszynski, 1930; Backlund, 1945; Bülow, 1957). Specific analyses of the vertical microdistribution indicate that both meio- and macrobenthic forms occur predominantly in strata well supplied with suitably decomposed organic particles, provided other abiotic conditions like oxygen are tolerable (Giere, 1970, 1971e). This became particularly clear where sand alternates with detritus layers (earlier surface horizons) (Giere, 197le, and unpubl. data). For tubificids, Pfannkuche (1977, 1980c) interpreted the accumulation of worms in subsurface horizons around the RPD-layer as a possible nutritive response to the rich bacterial stock developing in low-oxygen conditions (see p. 208). Turning from vertical to horizontal profiles, trophic specialization in combination with inhomogeneous distribution of food particles might contribute to the well-known patchiness of oligochaete occurrence even in apparently uniform areas (Giere, 1975) and, thus, would reflect only heterogeneous distribution of food, be it micro-organisms (Meadows & Anderson, 1968) or detritus particles with different bacterial colonization (Chua & Brinkhurst, 1973) and specific utilization (Wavre & Brinkhurst, 1971). More specifically, the distributional impact of food attracting oligochaetes has been demonstrated in experiments. Schulz (1955) who found baits of Fucus vesiculosus, buried in the sand, regularly frequented after 8–10 days by Enchytraeus albidus and Lumbricillus lineatus, assumed dissolved extracts washed out of the wrack layers by rain water to attract the worms from a greater distance. Similar attraction was reported by Tynen (1969) in experiments with L.reynoldsoni using Ascophyllum nodosum. He could

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prove decayed weeds to be significantly more attractive than the fresh thalli (see p. 208). Jansson (1968a) experimenting with meiobenthic oligochaetes, found pieces of Zostera in culture dishes to attract Marionina southerni and Aktedrilus monospermathecus (for details see p. 209). Giere (1975) and Hauschildt (1978) tested the distribution of Lumbricillus lineatus randomly inserted into Perspex columns which were filled with sterile moist sand and inoculated with thin layers of natural Fucus debris. After three days, 90% of all animals could be found feeding in the Fucus-horizon (Fig. 21).

Fig. 21.—Distribution of Lumbricillus lineatus in two sand columns after three days: control—in pure sand; experiment—in sand with Fucus detritus inserted in horizon 4–6 cm (modified after Hauschildt, 1978).

Predation on oligochaetes controlling their distribution has been only tentatively reported. In a North Sea beach, usually richly populated by interstitial oligochaetes, their absence in centres of turbellarian occurrences was related by Dörjes (1968) to intense grazing. From limnetic habitats, similar exclusions of oligochaetes have probably been evoked by predatory chironomid larvae (Brinkhurst & Kennedy, 1965; Loden, 1974). These results demonstrate that biotic factors, especially food, can attain, as well as other factors, a distributional impact of major importance for oligochaetes (Jansson, 1968a; Giere, 1975) and deserve far more attention in investigations on distributional patterns, usually interpreted in the light of the abiotic environment only. In these distribution analyses, experiments, although providing valuable information (see p. 259), cannot realistically simulate the integrated action of biotic and abiotic factors responsible for the natural distribution. Most suitable for field analyses are studies in the heterogeneous and fluctuating system of tidal shores, as exemplified in a vertical core from the supralittoral of a semi-sheltered North Sea flat (Fig. 22). The actual surface layer (0.5 cm thick) of fresh, not decayed Zostera leaves, directly exposed to light, evaporation or rainfall, and yet unattractive as food is almost avoided by the predominant enchytraeids

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Fig. 22.—Sediment structure, oxygen regime (recorded as Eh-values) and enchytraeid distribution in a vertical core from a sheltered North Sea sandy shore (Isle of Sylt; 1 m above high-water line) (modified after Giere, 1973).

Lumbricillus lineatus and Marionina spp. The well-sorted coarse sand mixed with

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Hydrobia shells (Md: 880 µm) of the following layer (0·5–2 cm), and only sparsely mixed with finer particles, indicates strong agitation by waves and renders this horizon, despite good oxygen supply (Eh-values above +300 mV), sufficient moisture, and reduced salinity fluctuations (around 20‰), unfavourable for oligochaete settlement. Decreasing grain size with an increasing fraction of fine particles and rich supply of aged detritus also in good oxygen conditions attracts most oligochaetes to the 2–5·5 cmhorizon. The almost complete absence of enchytraeids below these layers coincides with the sharp drop in oxygen content (Eh-values declining rapidly to oligoxic conditions around ±0 mV) in the fine to medium sand. Below 7.5 cm depth, the black sediment well mixed with muddy fractions, free of oxygen (Eh<–100 mV) and smelling of hydrogen sulphide, does not allow any colonization by oligochaetes. Whereas water content, pH, temperature, and salinity remained throughout the whole column in a range not limiting to oligochaete occurrence, the combination of strictly repellent factors (light, exposure to atmosphere, sediment agitation, lack of food and oxygen, presence of hydrogen sulphide) and favourable factors (aged detritus, rich oxygen supply in favourably sheltered horizons of suitable grain size) determined the highly heterogeneous vertical distribution of the worms.

TYPICAL ZONATION PATTERNS OF MARINE OLIGOCHAETA GENERAL OCCURRENCE Marine oligochaetes are distributed from the supralittoral down to the hadal zone (Fig. 23). The marine and brackish supralittoral is clearly dominated by

Fig. 23.—Sea depth distribution of some important oligochaetes (classification of depth zones after Vinogradova, 1970).

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enchytraeids, with the genera Enchytraeus, Lumbricillus, and Marionina predominating. In addition, several phallodriline tubificids are frequent in the mesopsammic intertidal zone, on boreal European shores mainly Aktedrilus monospermathecus, Phallodrilus prostatus, and Spiridion insigne, but they normally do not transgress the low-water line. Some members of the genera Marionina and Lumbricillus are, however, also abundant in the sublittoral, e.g. Marionina sublitoralis, described by Erséus (1976b) from 40–55 m depth, while the enchytraeid genus Grania is typically subtidal. Its deepest record so far is 500 m (G.macrochaeta pusilla, Erséus & Lassèrre, 1976). The widest range of occurrence is exhibited by phallodriline tubificids, a group of 12 genera (Erséus, 1980a) with representatives of the genus Phallodrilus occurring from the supralittoral down to the abyssal (P. profundus, Cook, 1970). Bathydrilus hadalis, closely related to Phallodrilus, was the first oligochaete described from the hadal Alëutian Trench (7298 m depth, Erséus, 1979a). The bulk of the species belonging to the genus Tubificoides inhabit the intertidal zone and the upper sublittoral, but some representatives from deeper waters are known, e.g. the abyssal T.aculeatus (Cook, 1969b). The ubiquitous species Tubifex costatus and Peloscolex benedeni are restricted to the intertidal and upper sublittoral. Their maximal depth recordings are 57 m for Tubifex costatus from Finnish coastal waters (Bagge & Ilus, 1973) and 52 m for Peloscolex benedeni in Cape Cod Bay (Cook, 1971). The naidid genera Amphichaeta and Paranais as well as Nais elinguis are limited to the intertidal and upper subtidal (<50 m depth); N. elinguis was encountered by Laakso (1969) in 34 m depth. SANDY BEACHES, WRACK ZONE The wrack zone is nearly exclusively inhabited by enchytraeids; characteristic species are Lumbricillus lineatus and Enchytraeus albidus. The distribution and abundance of enchytraeids depends on the location of the wrack on the beach, its composition and state of decay (see p. 208). Older wrack beds on the upper slope are dominated by E.albidus, often associated with some Fridericia spp. such as F.callosa (Michaelsen, 1927; Dürkop, 1934; Bülow, 1957), while the wrack zone on the lower shore is mainly inhabited by Lumbricillus lineatus. This species prefers more moist and more haline layers at the bottom of the wrack bed and is particularly frequent in the underlying detritus sands (Backlund, 1945; Schulz, 1955). These layers are also inhabited by a variety of Lumbricillus spp. (e.g. L.pagenstecheri, L. helgolandicus, L.rivalis, L.viridis, L.arenarius; see Pfannkuche, 1977) and Marionina spp. SANDY BEACHES, SUPRALITTORAL MESOPSAMMAL The enchytraeid genus Marionina, together with phallodriline tubificids, populates mainly the interstitial of marine and brackish-water sandy beaches. In addition, there occur some smaller Lumbricillus spp. (e.g. L. bülowi and L.knöllneri). Cernosvitoviella immota is frequent in the mesopsammal of Baltic beaches (Knöllner, 1935a; Bülow, 1957; Giere, 1976; Pfannkuche, 1980b). On European boreal shores, the tubificids Aktedrilus monospermathecus and Phallodrilus prostatus as well as the enchytraeids Marionina subterranea and M.spicula are the most common species in many types of

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beaches. In beaches of the central Baltic, M.southerni and M. preclitellochaeta are very abundant (Jansson, 1968a). Spiridion insigne has been reported from several western Baltic beaches (Knöllner, 1935a; Bülow, 1957; Pfannkuche, 1980b). Although all interstitial oligochaetes prefer well oxygenated sediment layers and avoid strata with oxygen deficiencies and formation of hydrogen sulphide (see p. 240), the different groups show a distinctive vertical distribution pattern (Giere, 1971e). McLachlan (1977b) describes a clear zonation pattern of “oligochaetes” (apparently phallodriline tubificids and enchytraeids of the Marionina group) which in mid- and low-tide areas occurred in the surface layers while at the high-water line and the top shore he encountered them in subsurface strata. If further differentiated, a marked divergence can be observed between tubificids and enchytraeids (Fig. 24), as was demonstrated by Pfannkuche (1980b). Enchytraeids prefer the upper strata in the sediment column beneath a dry surface layer with its great fluctuations of physiographical factors. Tubificids clearly become dominant at the bottom of the sediment column in the area of the ground water line which is characterized by relatively high water content and salinity as well as sufficient oxygen supply (Fig. 24).

Fig. 24.—Distribution of interstitial oligochaetes in a transect through a lotic beach of the western Baltic Sea: sample size, 5 cm×20 cm2; (modified after Pfannkuche, 1980b).

On a lotic western Baltic beach (Fig. 24), Marionina spp. (mainly M. subterranea) were most abundant in the middle and upper part of the slope with a maximum abundance 5– 10 m above the 0-m line (lowest point of the slope). Even at 15m, the animals were still present in some quantity. Below 4 m on the foreshore, Marionina spp. were, however, reduced in quantity. Vertically, a preference for the upper sediment layers is evident, maximum abundance was always found between 5–50 cm below the surface. These observations are in agreement with Jansson (1968a) from Swedish beaches. Lassèrre (1971b) found 85–90% of all enchytraeids at 10–70 cm depth and a maximum during

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seasonal peaks at 30 cm below the surface. In contrast to these Marionina spp., Aktedrilus monospermathecus aggregates on the low shore and in middle parts of the slope, but only close to the ground-water line; it was never found above the 10-m line. Maximum abundance was found between 3–6 m (Fig. 24). On a Baltic beach, this distribution pattern was observed in eight sampling profiles at different times of the year, A.monospermathecus always being close to the ground-water layer and even following its movements which occur irregularly due to hydrographic oscillations (Giere & Pfannkuche, 1978; Pfannkuche, 1980b). This basic distribution is in agreement with observations on this species by Fenchel, Jansson & Thun (1967), Jansson (1968a), Lassèrre (1971c), and Hulings (1974). Although exhibiting principally the same pattern, in a lenitic North Sea beach (Fig. 25) differences became apparent. In contrast to the rather steep Baltic slope, the North Sea shore rose from intertidal mudflats only populated by Tubifex costatus and Peloscolex benedeni (Giere, 1970). Due to this low inclination, the upper sediment layers of the low shore and even the central parts of the slope are far more moist than those of the Baltic beach. Again the animals concentrated in the moist layers of the low shore and in deeper strata of the upper slope, but here they were also found some distance above the ground-water line since oxygen deficiency around the ground-water layer forced them to move upwards (demonstrated by the RPDlayer in Fig. 25). These distributional differences could well be referred

Fig. 25.—Distribution of Aktedrilus monospermathecus in a sheltered North Sea beach (Isle of Sylt) in relation to redox discontinuity layer (RPD) and ground-water line (GWL); number in squares, individuals/5 cm depth ×20 cm2 (modified after Giere & Pfannkuche, 1978).

to preference and tolerance reactions (see p. 259). In our tests (Giere & Pfannkuche, 1978; Giere, 1977; 1980b) Aktedrilus was only rarely found in locations below 80% water saturation and its occurrence on the low shore and the central part of the slope is related to a salinity preference aligned with the specific habitat salinity (Pfannkuche,

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1980b). Tolerance experiments demonstrated its sensitivity to oxygen deficiency. A distinction of distribution patterns in species with overlapping habitat demands like the coexisting A.monospermathecus and Phallodrilus prostatus (cf. Pfannkuche, 1980b) is rather difficult and cannot be based only on reactions to physiographical factors without studying nutritional aspects (Brinkhurst et al., 1972). INTERTIDAL AREAS The composition of oligochaete fauna in the intertidal zone is much dependent on the sediment structure, as was demonstrated by Lassèrre (1967) and Giere (1970, 197le) for various meio- and macrobenthic species. On soft bottoms (mud, muddy sands) of European intertidal mud flats, macrobenthic tubificids like Peloscolex benedeni and Tubifex costatus, as well as the naidids Paranais litoralis and Amphichaeta sannio, predominate. A corresponding rôle is played by Tubificoides gabriellae in American muddy shores (Virnstein, 1979). Intertidal sands are inhabited by a variety of meiobenthic species mainly belonging to the phallodriline tubificids and the enchytraeid genera Marionina and Lumbricillus. Lassèrre (1971b), analysing the intertidal of the Bassin d’Arcachon, France, found marked differences in the distribution pattern of seven Marionina species. The populations of M.achaeta, M.weilli, M.mesopsamma, M.elongata, and M.spicula lived predominantly in the upper shore above mid-tide level. M.preclitellochaeta was most frequent near the mid-tide line, while M.subterranea prevailed between the high-water and low-water line. These differences in distribution were referred by Lassèrre to differing physiological adaptations to salinity, temperature, and oxygen regime, as was shown in experiments (Lassèrre, 1970, 1971b). SUBTIDAL HABITATS The occurrence and distribution of oligochaetes in the subtidal is mainly governed by water depth (Cook, 1971; Erséus, 1976a,b) and sediment structure (Cook, 1971, 1974; Erséus, 1976a; Erséus & Lassèrre, 1976; Pfannkuche, 1980c). In his comprehensive study on the Cape Cod Bay oligochaetes, Cook (1971) distinguished between a “fine sand and silt community” in the deeper parts of the Bay and a “coarse sand community” frequent in the shallower parts. Within these two assemblages, the relative abundance of a species was found to be controlled by the sediment composition (see p. 224). Peloscolex benedeni and Tubificoides longipenis coexisted over a wide range of coarse substratum, but maximum abundance of T.longipenis was recorded in shallower reaches (<15 m depth) while Peloscolex benedeni was encountered below 15 m depth. Cook (1971) suggested that Tubificoides longipenis was better able to exploit the finer sediments in the shallow stations. Erséus (1976a) studying the distribution of oligochaetes in Norwegian fjords, found T.amplivasatus and Crania macrochaeta pusilla on deep clay bottoms of 650 m and 400 m depth respectively. Bathydrilus rarisetis (depth range 20–400 m) occurred mainly in sand and gravel bottoms, whereas the enchytraeids G.variochaeta and G.maricola were frequent in sandy shell sediments only. Limnodriloides barnardi inhabited muddy shell

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sediments in 35 to 45 m depth while Peloscolex benedeni was encountered in soft bottoms above 25 m depth. Grania macrochaeta and G.postclitellochaeta were regarded by Erséus & Lassèrre (1976) as typical representatives of Amphioxus-sands. Temperature variations may also regulate the distribution of oligochaetes from the continental slope which probably are quite stenothermic (Erséus, 1976a). The factors controlling the distribution of bathyal and abyssal oligochaete species are completely unknown. The nearly immutable physical and chemical properties in the deep sea suggest that occurrence and distribution here is predominantly biologically controlled. CORALLINE AND CALCAREOUS SANDS The oligochaete fauna of these biotopes has been investigated quite recently and found to consist of a surprising variety of new species mainly belonging to the Phallodrilinae and to the enchytraeid genus Grania (Jamieson, 1977; Giere, 1979b; Erséus, 1980a). Occurring in considerable quantities, oligochaetes can be considered a significant component of the meiobenthos in coral reefs (Jamieson, 1977). Of particular interest are the numerous new forms which lack a functional intestinal tract and exhibit various degrees of gut degeneration. This trend can be interpreted as a nutritional response to specific chemical properties in these calcareous sands with probably highly enriched organic substances in the interstitial water. At least in the species, studied so far, bacteria, located in a space system between epidermis and cuticle (Giere, 1981; Richards, Fleming & Jamieson, 1981) seem to be implicated in the uptake of food. ESTUARIES AND BRACKISH INSHORE WATERS Our knowledge of the ecology of marine oligochaetes is still based largely on investigations of estuarine and brackish environments where both euryhaline marine and limnetic oligochaetes occur in considerable quantities. The first comprehensive works were from brackish areas of the western Baltic (Michaelsen, 1927; Knöllner, 1935a; Bülow, 1955, 1957), later to be followed and diversified by Jansson (1962, 1968a), Fenchel et al. (1967), Pfannkuche (1974, 1980a,c), Giere & Pfannkuche (1978). Brackish basins like the Baltic or the Black Sea and their adjacent bays have been preferred regions for important oligochaete studies (Dahl, 1960; Hrabě, 1966, 1971, 1973; Laakso, 1967, 1968, 1969; Finogenova, 1972; Bagge & Ilus, 1973; Moroz, 1974; Pfannkuche, 1980b). Other preferred locations are estuaries (Brinkhurst & Kennedy, 1962; Diaz, 1977, 1980; Hunter & Arthur, 1978; Birtwell & Arthur, 1980; Chapmann & Brinkhurst, 1980). In these studies, much attention has been given to the distribution of the various species along estuarine salinity gradients, as here salinity as well as sediment structure seems to be a crucial factor for faunal composition (see p. 236; Laakso, 1969; Moroz, 1977; Pfannkuche, 1980a,c). To demonstrate the distribution of oligochaetes along an estuarine gradient, the authors chose two regions, a brackish-water fjord (Schlei, western Baltic Sea) and a tidal estuary (River Elbe, opening into the North Sea, northern Germany). The Schlei-Förde (Fig. 26) represents for ≈40 km a salinity gradient from ≈3 to 16‰ S which corresponds in its

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range to a transect from the Kattegat region through the Baltic proper to the Gulf of Bothnia. There were 13 tubificid and nine naidid species encountered in the subtidal area of the Schlei-Förde. The oligohaline parts were mainly inhabited by eurytopic and euryhaline freshwater species (Stations A and B, Fig. 26). At Station A (annual mean 3.5‰ S), 95% of all tubificids are of limnetic origin; the only marine species, Tubifex costatus, reaches a mere 5.5% in relative abundance. The naidid fauna is represented altogether by nine species, seven freshwater forms, one typical estuarine (Amphichaeta sannio), and one truly marine species (Paranais litoralis) (Table XIV). At

Fig. 26.—Distribution and abundance of tubificids in the homoihaline SchleiFörde (Western Baltic) (modified after Pfannkuche, 1974, 1977).

TABLE XIV Distribution of naidids along the salinity gradient in the Schlei-Förde (modified after Pfannkuche, 1980c)

Salinity (‰) Species

2

4

6

8

10

12

Paranais lit1oralis

___________________________________________________________________

Amphichaeta sannio

___________________________________________________________________

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Nais elinguis

___________________________________________________________________

N.variabilis

__________________________

N.communis

__________________________

N.pseudobtusa __________________ Chaetogaster diaphanus

__________________

C.diastrophus __________________________ Stylaria lacustris

__________________________

Station B (5.1‰ S), Tubifex costatus becomes more abundant (30%), while the estuarine tubificid Monopylephorus rubroniveus occurs for the first time (34%). The remaining 36% are again represented by euryhaline limnetic species. The naidid fauna is here only composed of seven species. The silty sands of Station C (8.8‰ S), Station D (10.9‰ S), and Station E (11.7‰ S) are mainly inhabited by Tubifex costatus (>90%). Freshwater species are completely absent here, and only Monopylephorus rubroniveus occurs in some quantity (≈5%). The marine tubificids Tubifex pseudogaster, Clitellio arenarius, and Peloscolex benedeni are only found in small populations in the α-mesohaline part of the fjord (<5%) outside Station F, 15.6‰ S). The naidid fauna at Stations C, D, E, and F are only represented by Amphichaeta sannio, Paranais litoralis, and Nais elinguis. The relative and total abundance of tubificids was found to depend also on the sediment composition (see p. 236 and Fig. 15). In the α-mesohalinicum, sandy sediments exhibit the highest species diversity (5 species) and maximum abundance (40–60×103 ind.·m−2), whereas muddy sediments here were only populated by one species with<10×103 ind.·m−2. On the other hand, at the oligohaline stations, sandy substrata were populated only by<2×103 ind.·m−2 (nine species), while in muds the freshwater species developed much larger populations (cf. Brinkhurst & Kennedy, 1965; Pfannkuche, 1981). Regarding the zonation of these eurytopic species along the estuarine gradient, the granulometric impact is certainly overruled by the salinity regime, as nearly all species reported here can live in a variety of sediments (see p. 236). An interesting fact is the absence of Tubifex tubifex in the samples from the oligohalinicum, as this species was found very frequent in many surveys from tidal river estuaries (Palmer, 1968; Birtwell & Arthur, 1980; see below). It seems likely that it is substituted in the Schlei-Förde by Potamothrix hammoniensis which is also very eurytopic. Laakso (1969) also recorded only a few findings of Tubifex tubifex in coastal waters of the Gulf of Finland, whereas Potamothrix hammoniensis was common in many of his samples. Bagge & Ilus (1973) found P.hammoniensis exclusively in Finnish inshore waters. In addition to the above species in the Schlei-Förde, Nellen (1967) found in the same area the estuarine endemites Tubificoides heterochaetus and Monopylephorus irroratus, but omitted to give exact locations. The occurrence of tubificids in the poikilohaline Elbe Estuary (Fig. 27) differs in some

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respect from the Schlei-Förde when their distribution is based on the ‘normal’ hydrographical zonation. The tidal zone of the River Elbe extends for ≈145 km. The extension of the brackish-water zone is ≈55–60 km (Caspers, 1959), but can fluctuate considerably, as is demonstrated in Fig. 27. This unstable salinity regime restricts the occurrence of the marine species considerably more than in the more stable Schlei-Förde. Peloscolex benedeni, Tubifex pseudogaster, and Clitellio arenarius are found only in the outer polyhaline areas (Station A, Fig. 27), whereas Tubifex costatus is the only marine species really entering the Estuary (Stations D, E, F) together with the estuarine Monopylephorus rubroniveus (Station E), and is only abundant up to the border zone between the β-mesohalinicum and the oligohalinicum (Station F). (In homoihaline waters, the species is found down to 2‰ S, Laakso, 1969). The oligohaline section of the Elbe Estuary is only inhabited by freshwater species which also penetrate into the βmesohaline zone up to ≈8‰ S (Matthiae, 1977). When the possible fluctuations of the brackish-water zones are taken into account, as shown in Fig. 27, the distribution of tubificids coincides much better with the data from the Schlei-Förde. Then the marine species Peloscolex benedeni, Tubifex pseudogaster, and Clitellio arenarius are abundant in the α-mesohalinicum and Tubifex costatus even in the β-mesohalinicum. Limnodrilus hoffmeisteri is the dominant freshwater species in the bordering zone between the mesoand oligohalinicum (Station F). Tubifex tubifex becomes dominant from the βoligohalinicum followed by Limnodrilus claparedeanus and L.profundicola. In the freshwater tidal section, species diversity may increase considerably (Fig. 27), but is still reduced in comparison with other limnetic habitats.

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Fig. 27.—Distribution of tubificids in the tidal Elbe Estuary (Germany): lower section—salinity zones at low tide; upper section—salinity zones at high tide (combined after Caspers, 1959; Matthiae, 1977; Pfannkuche, unpubl.)

This estuarine succession of tubificids closely corresponds to the zonation in the Thames estuary (Hunter & Arthur, 1978; Birtwell & Arthur, 1980). Diaz (1980), comparing tidal estuaries from New England and Europe with lakes, related the diversity reduction of estuarine species to a lack of different habitats. One aspect of particular interest is the immigration of marine and estuarine species into the freshwater zone. Ax (1956) found the marine enchytraeid Marionina subterranea on a beach in the freshwater section in the River Elbe, ≈120km from its mouth. Aktedrilus monospermathecus was encountered by Graefe (pers. comm.) in the freshwater mesopsammal of the Elbe between the Stations I and K (Fig. 27). Amphichaeta sannio and Paranais litoralis were reported from the area of Station K (Pfannkuche, Jelinek & Hartwig, 1975). Occurrence of Aktedrilus monospermathecus and Spiridion insigne in phreatic waters was also described by Hrabě (1960) from the Weser Estuary (North Germany). Phallodrilus aquaedulcis, found at the same locality, has been considered by

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Cook & Hiltunen (1975) as a phallodriline freshwater species. Erséus (1979b), however, considered this species to be an estuarine endemite, as the water of the River Weser is highly anthropogenously chlorinated; Wachs (1963) reported a salinity of 0.7 to 16‰ from the River Werra, one of the main river sources of the Weser. He also found in these waters Paranais litoralis, Monopylephorus irroratus, and Lumbricillus rivalis. INLAND SALT WATERS The occurrence of these marine and estuarine species in isolated naturally salty inland waters (geogenic salts) has been reported by Boldt (1926), who found Monophylephorus irroratus and Lumbricillus lineatus in some north German inland brackish waters. A resampling at the same localities (Pfannkuche & Jelinek, unpubl.) added Paranais litoralis and Lumbricillus rivalis to these species Schmidt (1913) investigating inland brackish waters in Westfalia (Germany) found “L.lineatus” to be frequent in salinities from 39 to 62‰, which would document its amazing haline tolerance (see p. 260). He also mentioned the occurrence of this species in several south European salt ponds and springs.

TOLERANCE AND PREFERENCE REACTIONS IN MARINE OLIGOCHAETES Field studies, as previously mentioned, mainly advocate salinity, temperature, oxygen content, and moisture as factors determining the distribution of marine oligochaetes. There has been little experimental evidence on the basis of tolerance and preference tests, to decide to what extent single factors or combinations of factors are really operative in the interacting natural system, both with respect to local biotopes and to larger geographically disjunct areas. Differences between the ‘home range’ of populations and knowledge of the degree of ecophysiological modification would help to substantiate the often assumptive character of the complicated microdistribution on the basis of limitation, repellence and attraction within the web of environmental factors. The differences could also end up in consideration of the nature of the physiological disposition in these worms and of their adaptive potential regarding their geographical distribution. Most tolerance data available and commented on here deal with salinity and temperature, and occasionally with oxygen (see Tables XV and XVI). SALINITY TOLERANCE It is evident from Table XV that salinity tolerance of most larger marine oligochaetes, typically inhabiting the wrack beds of the seashore (e.g. the enchytraeids Enchytraeus albidus, Lumbricillus lineatus, L.reynoldsoni) or populating the shallow bottoms of tidal estuaries and bays (Tubifex costatus, Peloscolex benedeni, Paranais litoralis) covers an extremely wide range which underlines their good adaptation to the extreme and ever changing salinity regime in these biotopes. Even species which live preferably in freshwater habitats like the enchytraeid Lumbricillus rivalis, the tubificids Limnodrilus

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hoffmeisteri, Tubifex tubifex, and Ilyodrilus templetoni and the naidid Paranais frici tolerate amazingly high salinities, most exceeding their field range (Table XV). In the Firth of Forth, occurrence of Tubifex tubifex was

TABLE XV Experimental tolerance data of aquatic oligochaetes: arrows indicate: “up to”

Species

Salinity tolerance (‰ S)

Temperature tolerance (0°C)

Area of origin and remarks

Reference

Enchytraeus albidus Adults

0–70

W.Baltic, depending on Schöne, 1971 substratum

After gradual acclimation

→100

W.Baltic

Schulz, 1955

”

−11 to +34 ”

”

−13 to +36 W.Baltic, depending on Kähler, 1970 preacclimation

” Reproduction and development

0–200 0–40

”

”

Soil, Czechoslovakia

Krizenecky, 1916

W.Baltic

Schöne, 1971

1–25 Sewage beds, U.K.

Reynoldson, 1943

Embryos in cocoons

≈3–21

W.Baltic, depending on Schöne, 1971 season

E.barkudensis

2.5–30

Harbour mud, India

Subba Rao et al., 1980

W.Baltic

Giere, unpubl.

Lumbricillus lineatus

0–55

” L.reynoldsoni

2–36 East coast, U.S.A. 0–50

Irish Sea

Welch, 1914 Tynen, 1969

L.rivalis Adults

0–25

Ponlodrilus bermudensis

Sewage beds, Germany Giere, unpubl. 5–≈20 Sewage beds, U.K.

Development 5–25

Indian shores

Reynoldson, 1943 Ganapati & Subba Rao,

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1972; After acclimation

0–45

”

Subba Rao & Ganapati, 1975

0–34

Thames Estuary,

Birtwell & Arthur, 1980

Tubifex costatus At 5°C At 20°C

Species

0.5–34

(LT50-values: salinity: 7 days temperature: 4 →324 days)

Salinity Temperature tolerance (‰ tolerance (0°C) S)

Area of origin and remarks

Reference

Peloscolex benedeni At 5°C

2.8–34

At 20°C

5.9–34

”

”

”

”

”

”

Thames Estuary

Palmer, 1968

→28.5 Tubifex tubifex At 5°C

0–9.8

At 20°C

0–9.0 →33.9

Limnodrilus hoffmeisteri At 5°C

0–10.7

At 20°C

0–14.7 →37.5

Tubifex tubifex Non-adapted Adapted

0–7 →≈17

Limnodrilus hoffmeisteri Mature, acclimated to 0‰ S

→10

Immature, acclimated to 0‰ S

→11

” ” to 5‰ S

→14.5

(LT50:24h) Fraser River Estuary and environs of Victoria, B.C.

(LD50-values: 4 days, 10°C)

Chapman & Brinkhurst, 1980

Oceanography and marine biology

Tubificoides gabriellae

307

”

”

0.5–35?

”

”

Paranais litoralis

1–35

”

”

Monopylephorus rubroniveus

0–25

”

”

0→≈7.5

”

”

Mature acclimated to 10‰ S

1–30?

” ” to 20‰ S

2–30?

Paranais frici Acclimated to 15‰ S

Ilyodrilus templetoni

Species

Salinity tolerance (‰ S)

Temperature tolerance (0 C)

Area of origin and remarks

Reference

Psammoryctides barbatus

0–8

Dnieper-Lagoon, Moroz, 1974 Black Sea

Potamothrix hammoniensis

0–5

”

”

P.moldaviensis

0–5

”

”

Limnodrilus claparedeanus

0–5

”

”

Marionina southerni

0–40

E.Baltic

Elmgren, 1968

”

0–30

W.Baltic

Giere, unpubl.

M.subterranea

1.3–15

E.Baltic

Jansson, 1968a

M.preclitellochaeta

2.5–10

”

Jansson, 1962

M.spicula

3–30

30

French Atlantic coast

Lassèrre, 1970

M.achaeta

3–25

35

”

Lassèrre, 1971c

Aktedrilus monospermathecus

1.3–15

E.Baltic

Jansson, 1962, 1968a

Bermudian flat

Giere, 1977

1–35

A.monospermathecus LT50 5 h at 11‰ S

36

” 35‰ S ” 30°C ” 15°C

45

Interlinking of physical

” 11‰ S

35

” 35‰ S

32

” 15°C

43

” 30°C

35

782

Schlei-Förde, W.Baltic

Giere & Pfannkuche. 1978

limited by 4.1‰ S, and of Limnodrilus hoffmeisteri 7.7‰ S (McLusky, Teare & Phizacklea, 1980). An expanded ‘tolerance frame’, considerably wider than the usual occurrence would suggest, has also been reported by Moroz (1974) for tubificids from a freshwater lagoon with occasional influx of oligohaline water. The tolerance data for typical brackish-water forms (Monopylephorus irroratus, Marionina southerni) also exceed by far the range of their predominant occurrence: M.southerni enduring euhaline salinites well, is mainly restricted to the brackish beaches of the Baltic Sea (Jansson, 1968b; Giere, 1976). Tubificoides gabriellae, Tubifex costatus, and Paranais litoralis from coastal waters, where salinity fluctuations remain in the brackish range, can tolerate up to almost freshwater conditions (Tables XIII, XV). Salinity restriction is particularly rigorous for developmental stages (see Enchytraeus albidus in Table XV). Field data on Tubifex tubifex by Styczynska-Jurewicz (1972) show a decrease of salinity tolerance from 9‰ (adults) to 4‰ S for successful development. In contrast to this extremely wide “zone of resistance adaptation” (Vernberg & Vernberg, 1972), the experimental data given in Table XV for meiobenthic oligochaetes (mostly Marionina spp.) indicate restriction to the meso- or polyhaline brackish-water regime in the upper beach. TEMPERATURE TOLERANCE We certainly have to differentiate between survival, which is very little impaired, in unusually cold water and endurance in warm conditions. Rapid elimination by high temperature as a limiting element has already been pointed out by Reynoldson (1943) who related the semi-terrestrial occurrence of Enchytraeus albidus to the high tolerance range of the species, whereas restriction of Lumbricillus rivalis to the more aquatic microhabitats correspond with its less developed temperature resistance. In meiobenthic species, Lassèrre (1970, 1971c) found Marionina spicula to be limited to the lower beach by summer temperatures around 30°C, whereas M.achaeta with its higher temperature tolerance (35°C) could colonize the upper beach, and was well adapted to the low salinities recorded there. In contrast to this relatively clearly explained pattern from an area with warm summers, the tolerance ranges of Marionina spp. in studies from boreal beaches (Baltic Sea), done by Elmgren (1968), Jansson (1968a), and Giere (unpubl. data) cannot possibly be related to their horizontal zonation on the beach where M.southerni was found near the water-line, and M.subterranea 1 to 3 m higher up. Other ecofactors seem to co-determine occurrence, indicated perhaps by the results of Jansson (1966a) and Locy (in prep.) that M.subterranea possesses a marked preference

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for sediments of 200 to 250 µm grain size (see Table XVII). Interaction of several environmental factors, which enhance or attenuate adverse effects, could be proved in a comprehensive series of tolerance experiments to determine the interesting field distribution of the interstitial tubificid Aktedrilus monospermathecus from a small intertidal flat in Bermuda, Tucker’s Town Cove (Giere, 1977). In tests for temperature, salinity, pH, and oxygen thresholds and combinations of these factors, it could be shown that the resistance range of one single stress factor was amazingly wide (6–36°C; 5–40‰ S; up to 9.5 pH) where the oxygen supply was good and the concomitant factors stayed ‘normal’. Only the interaction of several stresses and/or oligoxic conditions caused considerable deterioration in viability (Fig. 3 in Giere, 1977). These reactions agreed well with the occurrence of the worms in the field; in the fine sand of the outer flat, high salinities (>35‰ S), often combined with extreme pH values (>9) and temperatures ≈35 °C prevailed at low tide in the surface layers. The worms could not avoid these rigid conditions by downward migration to the more attenuated subsurface milieu, because the sediment below 1 cm was anoxic. Therefore, here, the populations remained always suppressed (0–10 ind.·100 cm−3). In the slope of the beach around and above the highwater line the worms could, however, colonize in higher numbers (100–200 ind.·100 cm−3) the deeper strata with favourably reduced salinity, temperature, and alkalinity variations (33–34‰ S; 30°C; 8–8.5 pH) and a rich supply in food debris because the sediment was well oxygenated down to 10cm. This close adjustment of resistance capacity to oxygenated habitat conditions has also been found for polychaetes in experiments with decreasing oxygen tensions (Theede, 1973b; Theede, Schaudinn & Saffé, 1973). The physiological capacity of some species to survive even longer periods in hypoxic or anoxic water is emphasized in Table XVI especially for estuarine macrobenthic tubificids. The tolerance of Limnodrilus hoffmeisteri and Peloscolex benedeni mirror indirectly the ability to live in grossly polluted water (see p. 281). Compared with macrobenthic tubificids, the widespread Enchytraeus albidus seems to depend on a better oxygen supply. The physiological background for this difference is very poorly understood (see p. 273). Considering the extensive tolerance tests at various experimental conditions (Giere, 1977; Giere & Pfannkuche, 1978; Birtwell & Arthur, 1980), the additive stress effect of low oxygen and high temperature becomes obvious. In addition to experimentally supported explanations of complicated patterns of oligochaete distribution, these results lead to another aspect which made further studies on the widespread interstitial tubificid Aktedrilus monospermathecus highly interesting, namely comparison of the field distribution in allopatric and physiographically differing populations and their reactions in tolerance tests. This ‘zoogeographical’ method of interpreting tolerance results has already been considered for oligochaetes by Jansson (1968a), when he connected restriction of Marionina preclitellochaeta to the Baltic Sea with the limited salinity capacity of the species (5–10‰ S). The narrow salinity range (1– 15%) found by Jansson (1962) for Aktedrilus monospermathecus in tolerance experiments with populations from the eastern Baltic (see Table XV), may suggest geographical restriction to this brackish-water basin. Considering the wider temperature and salinity thresholds of this species from the highly saline Bermuda beaches as well as in the western Baltic and North Sea (Table XV), the general problem arose as to what

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extent tolerance ranges in separate populations of the same widespread oligochaete species can vary. Whereas the Bermuda worms always showed best survival in 35‰ S, regardless of temperature variations, the western Baltic populations survived best at 11‰ S (Fig. 28). For the North Sea worms (not shown in Fig. 28), survival rates were intermediate (e.g. 4 h at 35°C and 35‰ S). It is evident that the populations displayed best survival rates close to their average ‘home conditions’.

TABLE XVI Tolerance capacity to anaerobiosis in some tubificid and enchytraeids: temperature in ° C

Species

Duration of anaerobiosis (h)

Area of origin

Reference

Enchytraeus albidus

“Some hours”

Soil, Russia

Ivleva, 1960

”

8–16

Soil, Germany

Harnisch, 1942

Tubifex tubifex

4%-saturation: 500

British freshwaters Fox & Taylor, 1955

T.tubifex (LT50) 20°

28

” 25°

20

” 30°

13

Limnodrilus hoffmeisteri (LT50) 20°

52

” 25°

25

” 30°

18

Tubifex costatus (LT50) 20°

32

” 25°

17

” 30°

8

Peloscolex benedeni (LT50) 20°

59

” 25°

27

” 30°

18

Thames Estuary

Birtwell & Arthur, 1980

”

”

”

”

”

”

Bermudian flat

Giere, 1977

Aktedrilus monospermathecus (LT50:7% saturation) 7°

15

Oceanography and marine biology

” 15°

15

” 30°

15

” 35°

15

” 37.5°

8

” 39°

2.5

” 40°

0.1

” 7°

24

” 15°

24

” 30°

26

” 35°

2

” 37.5°

1

311

Schlei-Förde, W.Baltic

Giere & Pfannkuche, 1978

” 40°

0.3

Tubifex tubifex

600

Swedish fresh waters

Alsterberg, 1922

Limnodrilus hoffmeisteri

”

”

”

0–2°

1000

Dausend, 1931

18–20°

216

German fresh water

“Tubificids”

For temperature the same is true although, under normal salinity conditions, survival at 35°C—the home range of the Bermuda worms—is only slightly better than that of the boreal populations. Introducing an additional stress, however, like low oxygen, causes differences in survival times between the two populations in question (>15 h compared with 2 h; see Table XV). The above results underline again the governing rôle of oxygen supply for (meiobenthic) oligochaete survival even in species of a considerable

Interlinking of physical

786

Fig. 28.—Salinity tolerance of Aktedrilus monospermathecus (Baltic and Bermudian populations): survival time (LT50) at various temperatures (from Giere & Pfannkuche, 1978).

euryoecious nature. The oxygen regime not only determined the colonization pattern in the Bermuda flat, but also that in the North Sea shore of Keitum (Fig. 25), where anoxic conditions (indicated by the RPD-layer) were established at about the same depth as the ground-water layer and forced the population to move upwards with concentrations always above ground water. It is still not fully understood, however, which factors deter this species from populating also the upper horizons, especially at the top shore, which offer both good oxygen, salinity, and temperature conditions. In a temperate climate, the tolerance values obtained defined a wider “zone of resistance” (Arndt, 1973) than realized by natural habitat conditions. Here the choice of preference zones, as a general ecological principle narrower than the tolerance limits (Kinne, 1970, 1971), allows a more differentiated analysis of the animals’ distribution; they orientate themselves along horizontal and vertical gradients of ecofactors (Gray,

Oceanography and marine biology

313

1965; Jansson, 1967b; Tynen, 1969). PREFERENCE EXPERIMENTS The need for preference experiments, in which the animals could adjust their microdistribution in gradient fields, more comparable with the situation in nature, has already been mentioned by Jansson (1967a) and Kinne (1970, 1971). It is symptomatic, however, for the little developed ‘state of the arts’ that only few and somewhat isolated tests on marine oligochaetes have, so far, been performed. Results in Table XVII show that the response of the

TABLE XVII Experimental preference data of aquatic oligochaetes

Species

Preference

Area of origin

Reference

Enchytraeus albidus

5–1 5‰ S for development, W.Baltic Sea reproduction

Schöne, 1971

”

3–10‰ S

Baltic Sea

Backlund, 1945

”

no marked temperature preference

W.Baltic Sea

Schulz, 1955

”

5–15‰ saturation: moisture preference

”

”

”

19 °C

Terrestrial Krüger, 1955 populations, Germany

E.barkudensis

Around 25‰ S

Indian harbour, bottom Subba Rao et al., 1980

Lumbricillus lineatus

no salinity preference

Irish Sea

”

no salinity preference

Baltic Sea

Backlund, 1945

L.reynoldsoni

15‰ S

Irish Sea

Tynen, 1969

Marionina subterranea

16–20 °C

Pacific coast of California

Locy (in prep.)

”

5‰ S

E.Baltic Sea

Jansson, 1968a

M.southerni

1–5‰ S

”

Elmgren, 1968

M.preclitellochaeta

0.3‰ S

”

Jansson, 1962

Aktedrilus monospermathecus

2.5–5‰ S

”

”

Marionina subterranea

125–500 µm sediment

”

Jansson, 1966a

”

200–250 µm sediment

Pacific coast of

Locy (in prep.)

Tynen, 1969

Interlinking of physical

788

California Tubifex costatus

250–1000 µm sediment

Thames Estuary

Birtwell & Arthur, 1980

Peloscolex benedeni

63 µm sediment

”

”

Tubifex tubifex

63 µm sediment

”

”

Limnodrilus hoffmeisteri 63 µm sediment

”

”

worms tested usually matches their habitat conditions fairly well (see also Jansson, 1968a,b). Again, the considerable physiographical accommodation of some typical wrack bed forms to extremes of salinity and temperature is striking. We also started our preference studies with these larger ubiquitous species, but focused later on the interstitial Marionina southerni and M.subterranea and, particularly, on Aktedrilus monospermathecus in order to further clarify their different local distribution pattern. By choosing species with disjunct populations, comparing our own results with those of other authors, and testing pre-acclimated animals, we wanted information about the diversity and flexibility of preference responses in some common marine oligochaetes. The scarcity of preference studies may partly be due to problems in experimental set-up. So, our experiments were preceded by the development of reliable and versatile devices for testing preference in salinity, temperature, and moisture gradients (Giere, 1979c). The salinity device was principally based on a series of Jansson’s (1962) “alternative chamber”. Temperature gradients in moist sand were maintained in a trough adequately temperature-regulated by thermostats. Moisture gradients were set up in an oblique, sandfilled trough supplied with sufficient water at one end to provide a complete range from saturation to dryness. All devices had to meet the requirement of smooth and well replicable gradients without abrupt “bordering zones”. Considering the large mobility even of small interstitial oligochaetes (up to 3 cm·min−1, according to Jansson, 1962), a duration of 48 h per experimental run should be adequate for orientation of the worms. The results selected here are based on several replicates and careful statistical treatment (Giere, 1980b). Our salinity experiments with Lumbricillus lineatus populations from the Baltic (Fig. 29) showed a clear preference for the low ‘domestic’ salinity,

Oceanography and marine biology

315

Fig. 29.—Salinity and temperature preference of Lumbricillus lineatus: comparison of North Sea and Baltic populations; in brackets, average habitat salinity; (from Giere, 1980b).

whereas worms from the North Sea displayed a much more diffuse pattern, although a trend towards low salinities was recognizable. This is in contrast to the lack of any preference in the experiments of Backlund (1945) and Tynen (1969) (see Table XVII). The use of an artificial substratum in their devices may contribute to this contradiction (Giere, 1979c) as well as some possible variance in genetical background (see below). Tynen stressed the correspondence between littoral zonation and response to salinity for L. reynoldsoni, a species from the upper shore with low salinities, and for L. lineatus which lacked any marked salinity preference and duly was distributed all over the shore. The temperature preference of L.lineatus (Fig. 29, upper curve) reflects well the mean habitat conditions in their temperate climate. Long-term adaptation (>30 days; 26°C), however, shifted the preference significantly towards a higher level. Conditions were more complicated in Enchytraeus albidus, the other dominant wrack form (Fig. 30). First salinity tests in summer 1979 with

Interlinking of physical

790

Fig. 30.—Salinity preference of two allopatric Enchytraeus albidus populations: comparison of preference responses in non-adapted and long-term adapted worms (from current experiments).

North Sea populations showed an almost homogeneous distribution along the gradient from about 2 to 55‰ S (not depicted in Fig. 30). In a later series in spring 1980 a clear preference for salinities between 15 and 35‰ became apparent (Fig. 30), which agrees with habitat conditions in the North Sea shore. This range was maintained by the worms even after a 4-wk adaptation period prior to the test runs in moist sediment gradually diluted to freshwater conditions. Our salinity tests with Baltic populations, not shown in Figure 30, resulted in a marked preference for salinities below 10‰, and particularly below 5‰, which was confirmed by corresponding values in experiments with Baltic worms performed by Backlund (1945) and Schöne (1971). Here, the stability of this preference range after pre-adaptation to euhaline salinities has not yet been tested by us. Corresponding tests with conspecific populations from garden soil consistently exhibited a marked preference for oligo- and mesohaline conditions (0–15‰ S), which also remained stable even after a 4-week preacclimation of the worms to sea water of 17‰ S (Fig. 30). This consistency, contradictory to conditions in Lumbricillus lineatus, was seen also in temperature-choice experiments both with terrestrial and North Sea populations (preferred range about 12–20°C; this corresponded to room temperature

Oceanography and marine biology

317

during maintenance). There is some evidence that results in Enchytraeus albidus coincided, when collected on identical sampling dates, i.e. worms taken in summer 1979 showed little preference behaviour for salinity, whereas the early spring population showed a clear salinity choice. This would imply that the physiological status, varying with the seasons, strongly influences salinity responses. Similarly, young E.albidus, studied by Kähler (1970), had a higher mortality in spring, if exposed to high salinities, than in autumn. The ecological implication of this differing preference response in E.albidus is as yet uncertain. A seasonal variability in salinity choice was also recorded for other marine invertebrates by Theede (1973a), Percy (1975), and Bodoy, Dinet, Massé & Nodot (1977). Flexibility of the salinity adjustment after acclimation was monitored for the euryhaline tubificids Limnodrilus hoffmeisteri and Tubificoides gabriellae (Chapman & Brinkhurst, 1980; see Table XV). In the interstitial enchytraeid Marionina southerni, agreement between habitat conditions and experimental results was found in temperature (Fig. 31, upper curve) but in salinity, the western Baltic populations with

Fig. 31.—Salinity and temperature preference of Marionina southerni from the Schlei-Förde (Western Baltic) (from Giere, 1980b).

habitat water of ≈10‰ S were always aggregated around 5‰ S. Close agreement to similar results by Elmgren (1968) with populations from the eastern Baltic (local salinity ≈6‰) might lead to the conclusion that this species has a stable, fixed salinity preference. By long-term adaptation, however, the preference peaks became somewhat reduced with a clear trend towards higher salinities. We also tested the temperature preference of Marionina subterranea from the North Sea, for comparison with corresponding experiments carried out by Locy (in prep.) who used populations from the Pacific coast of California. Both populations proved to have a

Interlinking of physical

792

well defined, and not significantly different temperature preference. On the Pacific coast the preference was ≈16–20°C, and in the North Sea ≈15–16°C, which could be slightly shifted after a 5-wk acclimation in 3°C to 12–14°C. For Aktedrilus monospermathecus, we had, as in the earlier tolerance tests, the chance to study allopatric populations from physiographically differing habitats (Fig. 32). Again, the salinity preference agreed fairly well with the

Fig. 32.—Salinity and temperature preference of Aktedrilus monospermathecus: comparison of Baltic, North Sea, and Bermuda populations (from Giere, 1980b).

‘home conditions’. The North Sea animals reflected the wide tidal plus non-tidal salinity fluctuations in that shore line. Only the Baltic populations exhibited a certain discrepancy between average salinity (≈11‰) and preference data. Also temperature ranges could be related to prevailing biotope conditions. In the Bermuda populations, the preference range explains the sub-optimal conditions in the surface layers of the sand flat far clearer than did tolerance data. The preferred attenuated temperatures between 25 and 30 °C and salinities of 35‰ were typical for the deeper layers of the high-water zone and the upper shore. It remained to be understood why A.monospermathecus, if occurring in the top shore, was restricted to near ground-water layers. As in this ubiquitous species, no substratum specificity seems to exist (judging from sediment conditions at numerous sampling sites and experiments reported by Jansson, 1968b), our studies focused on moisture gradients as a causative factor. First tests showed a marked preference for very high water saturation (>90% of the inserted worms preferred saturation between 80 and 100%) thus strongly suggesting a well-developed response to moisture differences. A similar

Oceanography and marine biology

319

distributional problem, probably to be explained by different moisture demands, is the consistent occurrence of Lumbricillus lineatus close to the water-line compared with Enchytraeus albidus in the upper shore. Preference experiments on this complex are still in progress (Giere, unpubl.). Comparing the tolerance and preference results presented above, it becomes evident that tolerance data often correspond more to ecological extremes and rather represent ‘momentary situations’, particularly as they are generally found in unnatural single factor stress conditions. As a general rule, the regular conditions for existence with successful reproduction and embryogenesis are better described by the preference range. Especially in the subtropical Aktedrilus population, the tolerance thresholds closely matched the extremes of salinity, temperature, pH, and oxygen supply at the sand surface. This close alignment between field distribution and tolerance capacity in warmer climates has already been stressed for other animal groups (Jansson, 1966b; Alderdice, 1972; Vernberg, 1975; Wieser, 1975). In a temperate climate, the short-term tolerance thresholds of most euryoecious oligochaetes were found to exceed the range of physiographical factors (exceptions were anoxia and presence of hydrogen sulphide). Hence, in boreal zones their remarkably discontinuous distribution reflects more the preference levels than tolerance limits, but interference with other kinds of preferences, with biotic factors and long-term (seasonal) variations often blurs the situation. For instance, substratum selectivity certainly plays a more important rôle (see p. 224) than the few existing preference results (Table XVII) would suggest. Interaction of all these factors may explain the differences between onefactor tests (salinity, temperature) and actual field situations as found in some of our experiments (e.g. Aktedrilus from the western Baltic—salinity preference lower than the averaged local salinity). The impact of pre-acclimation to shift the preference ranges of many macrofauna and meiofauna species, as shown in the literature and in our experiments, seems to be a general feature in eulittoral oligochaetes. Some conflicting results, particularly with Enchytraeus albidus (Backlund, 1945; pers. obs.) have to be further analysed. The data of Elmgren (1968) and Jansson (1968a) on Aktedrilus monspermathecus, indicating a stable preference irrespective of the background conditions, must probably be assigned to the relatively insignificant biotopic differences between the test populations or an insufficient acclimation time (20 days; 20‰ S) considering the extremely euryoecious nature of this species. The flexibility of tolerance and preference ranges throws some light on the nature of adaptive responses in these oligochaetes. In most cases they seem to represent nongenetic adaptations (Kinne, 1964, 1971) accomplished by exposure to different ecological conditions over longer periods, and particularly stable as long-term modifications after acclimation of the sensitive juvenile stages (Percy, 1975). These steady-state reactions (Kinne, 1970) probably are superimposed on a wide genetically fixed physiological background which may even keep covered conditions never to be encountered by a local population in its particular environment (e.g. low temperatures in the Bermuda population of Aktedrilus). To generalize on this interpretation may yet be premature considering the widely occurring genetic inter-population variance in marine littoral species (see compilations by

Interlinking of physical

794

Battaglia & Beardmore, 1978; for polychaetes by Grassle, 1980). In oligochaetes, Christensen and his school were the first to draw attention to intraspecific genetic variance in Lumbricillus lineatus and L.rivalis from Danish regions, not only through alloploid caryotypes (p. 191; Christensen & Nielsen, 1955; Christensen & O’Connor, 1958; Christensen & Jensen, 1964; Christensen, Jelnes & Berg, 1978), but also evidenced as enzymatic differences in sibling species. Christensen & Jelnes (1976) proposed to correlate this variance in the electrophoresis pattern of the different populations in L.rivalis to the salinity regime in the respective habitats on the shore, reasoning that the population from the seashore represented the original ecological adaptation. Christensen, Berg & Jelnes (1976), discussing mechanisms that maintain genetic polymorphism in mixed populations of diploid and triploid L.lineatus, hypothesized that in a heterogeneous environment like brackish shores different genotypes would be favoured in occupying different ecological niches. In our own tentative electrophoretic studies (unpubl.), L.lineatus populations from Baltic and North Sea shores showed major differences in band position. If similar results with reproductively isolated Enchytraeus albidus samples (North Sea, Baltic Sea, garden soil) can be substantiated in further studies, genetic variability could well represent another reason for the inconsistent preference responses, and also for some discrepancies in tests with this species using hydrocarbons (see p. 291). Electrophoretic studies, scrutinizing the ‘anchoring’ of divergent ecological responses, examining the possible establishment of physiological races or the evolution of sibling species are especially needed in littoral oligochaetes, as cross-breeding experiments can hardly be performed with these small hermaphrodites.

PHYSIOLOGICAL MECHANISMS AND ADAPTATION Within the framework of this review, compilation of physiological data will be restricted to aspects of ecological relevance, i.e. to osmoregulative processes involved in salinity and temperature tolerance, or to metabolic pathways determining respiratory capacities. The scarcity of physiological data on marine oligochaetes has been impressively documented (Mill, 1978). Limnetic forms like Tubifex tubifex and Limnodrilus sp. are discussed in some detail, but marine species hardly mentioned. Hence, some of the results reported here will be supplemented by data from limnetic oligochaetes and their better investigated polychaete relatives. Physiological studies on aquatic oligochaetes referring to an ecological background focus on three main aspects: control of body volume and osmoregulation of body fluids; respiration, often in relation to osmoregulation, as a response to altered environmental conditions, and physiological adaptations and mechanisms involved in temperature changes. Eulittoral oligochaetes, adapted to a habitat with heavily fluctuating salinity, exhibit an effective regulation of body volume, closely connected with a control of ionic concentration in the coelomic fluid. This seems to refer both to small mesopsammic and to large mud-inhabiting species. Lassèrre (1975b), summarizing his detailed studies on Marionina achaeta, divided the process of volume regulation in lowered salinities into a short-term response to excrete water rapidly from the permanently hyperosmotic body (≈ 2–4 h) and a subsequent longer lasting regulatory phase (5–15 days) during which the

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body weight and volume was gradually re-adjusted to initial values. Restitution of the initial status by active osmoregulation was only possible between salinities of 3 and 25‰. Beyond these thresholds (from 0.5–3‰ and from 25–35‰ S), the effectiveness of regulation decreased or broke down completely (fresh or hypersaline water) leading to complete osmoconformity. Maintenance of slightly hyperionic body fluids by an effective osmoregulation within a certain, specific “span of compensation” with short adaptation times (common in many euryhaline polychaetes like Nereis spp.) seems to be a general feature in marine oligochaetes. Enchytraeus albidus regulates its internal ionic concentration over a wide range becoming an osmotic conformer only in very high salinities (Drawert, 1968; Schöne, 1971). Pontodrilus bermudensis is hyper-regulating too, but only in salinities <15‰, this phase commencing at ≈200 mM NaCl·1−1 and ending at >15‰ (Subba Rao, 1978a,b). Even freshwater species like Tubifex tubifex keep their haemolymph concentration constant at 4‰ S between 0 and 6‰ S of the surrounding medium (Styczynska-Jurewicz, 1972). Only the littoral Lumbricillus reynoldsoni is reported by Tynen (1969) to be poikilosmotic. Lack of any ability to osmoregulate would, however, be surprising in secondary marine animals of freshwater origin as oligochaetes are held to be. It is possible that for L.reynoldsoni the concentration tested (full sea water) was already beyond the “span of compensation”. More detailed experiments with this species could provide the answer to this question. It was clearly shown by Lassèrre & Renaud-Mornant (1973) and Lassèrre (1975b, 1976) that there exists a close relation between the range of osmoregulation and habitat conditions of the respective species, a principle generalized by Oglesby (1978) for all annelids. The physiological mechanisms involved in this osmoregulation seem to include an active ionic transport using the Na+/K+ -ionic pump, whereas there is as yet little, if any indication for an active Cl− -transport through the epidermis (Lassèrre, 1975b), so important for osmoregulation in nereids (Jørgensen & Dales, 1957). Lassèrre (1976) reasons that a substantial part of the oxygen consumption and of the energy gained as ATP from respiratory metabolism is utilized for the Na+/K+ -pump mediated by a hydrolysing ATPase system which is, therefore, involved in active ionic regulation. In a diluted medium, re-absorption of Na+ through the nephridial walls results in a hypotonic urine excreted through the nephridial pores. In more saline water, attempts to maintain hyperosmotic body fluids via Na+-absorption are supported by permeability changes of the epidermis and possibly also of the intestinal wall (Lassèrre, 1975b). The regulatory impact of specific ions such as Na+, K+, and Ca2+ is emphasized in experiments with Enchytraeus albidus (Kähler, 1970) in which a change in the ionic balance modified the physiological disposition and the salinity-temperature range of the worms. Removal of Na+ from the highly saline external medium resulted in a 30% decrease in oxygen consumption (Lassèrre, 1969, 1975b). Judging from results with Lumbricus terrestris (Ramsay, 1949a,b), this regulation of urine concentration also includes permeability changes of the nephridial tube walls. Lassèrre (1975b) stressed the “energetic capacity” of the nephridial walls with their numerous mitochondria and cellular partitioning by double membranes. In Enchytraeus albidus, Richter & Gersch (1967) found that alteration in the ionic composition of body fluids markedly changed the membrane potentials of neurosecretory “P”- and “Q”-cells of the brain. Drawert (1968) also proved the existence of a close

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connection between ionic concentration of the ambient medium, osmoregulation, and neurosecretion in the same species. Specimens from garden soil showed, after exposure to sea water, rich accumulation of secretion in the “Q-cells” located posterio-marginally in the cerebral ganglion, and in “U-cells” of the ventral nerve-cord. Deletion of the “Qcells” resulted in an increase of body-wall permeability. Drawert stated that some neurohormones may regulate both the sodium concentration in the coelomic fluid and the permeability of body epithelia. He also suggested that production of mucus is directly involved in the active ionic transport judging from the reaction of mucous cells in the nerve cord of E.albidus and Lumbricillus lineatus after keeping the worms in sea water. Recent experiments by Ferraris & Schmidt-Nielsen (1979) with Clitellio arenarius corroborate this cerebral and neurosecretory influence on osmoregulation and nephridial function in oligochaetes. After some hours in 50% sea water, the Na+ and Cl− contents (but not K+) were significantly higher in decerebrated worms than in the controls. Decrease of the ability to eliminate excess of extracellular fluid was indicated by the higher water content in the ablated worms. Subba Rao & Ganapati (1975) found in Pontodrilus bermudensis, that Cl− and K+ were hyperionic and reasoned that free amino acids in the body fluid played a rôle as a compensating means for maintaining osmotic equilibrium in hypersaline sea water (see also Subba Rao, 1980). Whereas the sensory basis for this apparent ability to discriminate between changes in environmental conditions is completely unknown (Oglesby, 1978), the high energy demand of the regulatory reactions has been repeatedly recorded. Oglesby (1978) quotes an increase in oxygen consumption of between 30 and 100% in low salinity compared with “normal” rates. Lassèrre (1969) calculated in Marionina achaeta that 20–30% of the rate of oxygen uptake ( ) was used for the various metabolic compensations and he, therefore, took respiration rates as an index of homeostatic capacity (Lassèrre, 1969, 1970, 197 le, 1975b, 1976). He monitored a marked increase in as an immediate response to transfer into a markedly different salinity and a subsequent decline as adaptation proceeded. Both for M.achaeta and M.spicula, Lassèrre recorded, however, no permanent change in

over a wide salinity

range. The same holds valid for in a wide range of ambient oxygen tensions (Lassèrre & Renaud-Mornant, 1973). Only beyond the “adaptive plateaus” of 3–25‰ S (M.achaeta) and 5–30‰ S (M.spicula) did increase considerably (Fig. 33a). Lassèrre related his respiratory recordings to the ecological background of the species studied. Characterizing M. achaeta and M.spicula as “effective euryoxybiontic regulators” he stressed their adaptation to life in a beach with frequent oxygen deficiencies and heavy fluctuations in salinity. He even could relate his physiological results to the horizontal distribution of the two species with M.achaeta in the top shore and M.spicula closer to the water-line (see p. 254). Palmer (1968) recorded a comparatively stable over the entire salinity range tolerated by Tubifex tubifex after long adaptation periods (0–7‰ S; Table XVIII). The interrelationship between metabolic effort and external temperature

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Fig. 33.—Respiration rates of two interstitial oligochaetes: a, respiration at various salinities; b, respiration at various temperatures; comparisons based on mean oxygen consumption adjusted to a mean wet wt of 30 µg; DW, freshwater; TW, water with Cl− added; micro-respirometer with standard Cartesian diver (modified after Lassèrre, 1971c).

seems to correspond to the salinity situation (Lassèrre, 1975b, 1976). For both Marionina spp. Lassèrre recorded an adaptive plateau with a relatively constant oxygen consumption of 1×10−3 µl O2·µg−1·h−1 between 10° and 28°C (Fig. 33b). Below 10°C, the respiration decreased for both species; above 28°C, m.achaeta, living in the upper and often

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considerably ‘heated’ shore could maintain its consumption up to 35 °C while for M.spicula increased rapidly. A rather wide compensation for the effect of temperature on (anaerobic) metabolism (between 15° and 25 °C) has been reported for Tubifex tubifex (Brandt, 1978). The lability of this regularity capacity and the possible interaction of concomitant environmental factors is shown by the breakdown of temperature compensation after starvation.

TABLE XVIII Respiration rates of selected oligochaetes

Species

O2-consumption (µl·mg wet wt−1·h−1)

Experimental conditions

Reference

Marionina achaeta

0.76

15‰ S; 20°C; base don Lassèrre, 1969, mean wet wt of 30 µg 1971c, 1976

M.spicula

0.9

”

Enchytraeus albidus

0.48

20°C; 10 mg live wt

O’Connor, 1967

Fridericia galba 0.15

20°C; ”

Lumbricillus lineatus

0.54

20°C; ”

L.rivalis

0.45

20°C; 9–10 ”

L.viridis

0.45

20°C; 13 ”

L.lineatus

(derived from dry wt data):

Mature worms, marine Regier,* (unpubl.) cultures from North Sea:

0.14

5°C; 15‰ S

0.2

10°C; 15‰ S

0.42

15°C; 15‰ S

0.4

20°C; 15‰ S

no consistent values, but rapid 25°C; 15‰ S increase of O2-consumption and high mortality Mature worms, Baltic Sea:

L.rivalis

0.16

5°C; 15‰ S

0.17

10°C; 15‰ S

0.87

20°C; 0‰ S

Williams et al., 1969

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0.41

10°C; 0‰ S

”

Enchytraeus coronatus

“Similar values”

”

”

E.albidus

0.35

19°C; based on 150mg Krüger, 1955 samples

Lumbricillus “lineatus”**

0.33

” from sewage water

Enchytraeus albidus

Based on 450 mg Ivleva, 1960 samples, from cultures in soil 0.1

Recorded in water

0.4

Recorded in air

Achaeta eiseni

0.3

Based on 160 µg·worm−1

Fridericia bisetosa

0.3

Based on 440 µg·worm−1

Tubifex tubifex

0.57

20°C; 0‰ S

0.58

20°C; 7‰ S based on 2.5 mg·worm−1

0.31

25°C

0.36

25°C

Limnodrilus hoffmei

”

Nielsen, 1961, (recalculated by O’Connor, 1967)

Palmer, 1968

Van Hoven, 1975

teri

Ilyodrilus templetoni

* We thank Mr M.Regier, Hamburg, for assistance in compiling this table. ** Probably wrong identification for Lumbricillus rivalis.

Judging from polychaete studies, it is possible that there exist specific physiological adaptations to extreme temperatures causing seasonal and geographical climatic adjustment for populations from widespread species. This adaptation seems to be mediated by a neurohormonal ‘factor’ released from the central nervous system (Mangum, 1978). Weber (1978) considers oligochaete “erythrocruorins” with their high oxygen affinity as superior to extracellular haemoglobins. In addition, the very low “oxygenation heat” found in Tubifex as in some euryoecious polychaetes could be of adaptive value for species in intertidal flats, maintaining a considerable oxygen affinity even at high temperatures. Under these conditions, the negative impact of diminishing oxygen affinity will be compensated, in the presence of a Bohr effect, by the concomitantly decreasing pH-values of the blood. The general increase of respiration rates with rising temperatures (until a specific upper threshold), apparently follows different lines in the oligochaetes studied, if experimental

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methods are reliable. Whereas Lassèrre (1971c) found an “adaptive plateau” of fairly constant values, well compensated for temperature elevation for marine Marionina species (Fig. 33b), O’Connor (1967) reported for both congeneric terrestrial species and other eulittoral enchytraeids a continuously rising oxygen uptake when exposed to higher temperatures (Fig. 34). Hence, attempts to draw general conclusions in calculating a uniform temperature-related multiplication factor of respiration, as stated by Williams, Solbé & Edwards (1969) for Lumbricillus rivalis ( 10°C) or by

at 20°C 1.62 times higher than at

Fig. 34.—Oxygen uptake of some terrestrial and aquatic enchytraeids in relation to temperature (modified after O’Connor, 1967).

Brazda & Rice (1940) for Tubifex tubifex (at 30 °C 1.3 times higher than at 25 °C) have to be interpreted with caution regarding the specific variance and inconsistency in the slope of respiration curves: in Marionina achaeta,

increased between 2° and 5°C

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about 10 times, whereas uptake was kept constant between 10° and 20 °C (Lassèrre, 1971c). Similarly, Tubifex tubifex shows an extreme regulatory capacity for increasing salinity (Palmer, 1968) which keeps respiration almost constant between 0 and 7‰ S. In Lumbricillus lineatus, increases, according to Regier (Table XVIII), in North Sea populations from 5° to 10°C, 1.4 times, from 10° to 15°C, 2.1 times, but remains fairly constant from 15° to 20°; in Baltic populations, almost no increase could be recorded between 5° and 10°C. Warwick (1980) related temperature-induced -Patterns (in nematodes) to food supply, reasoning that species with stable food conditions possess “adaptive plateaus” in contrast to worms with temporally changing availability of their diet. From intergeneric comparisons with enchytraeids of varying size, it emerges that respiration rates decrease with increasing body weight (Fig. 35; Krüger, 1955; O’Connor, 1967). The latter author also showed that

Fig. 35.—The relationship between weight and oxygen uptake in several marine enchytraeid species (modified after O’Connor, 1967).

littoral oligochaetes such as Lumbricillus and, to a lesser extent, Enchytraeus living in relatively wet habitats, have a respiration rate three times higher than Fridericia spp. of comparable weight from dry soils (see Table XVIII). He interpreted this phenomenon as adaptation to increasingly terrestrial conditions where reduction of water loss through thicker body walls coincided with low respiration activity. This argument is, however, not corroborated by the data of Krüger (1955) on the terrestrial Enchytraeus albidus compared with the aquatic Lumbricillus “lineatus”. Although comparison with respiration data from other authors are impaired by different experimental procedures and calibrating references, the values compiled in Table XVIII support the inverse relation between body weight and oxygen uptake emphasized by Krüger (1955) and O’Connor (1967). Hence, interstitial oligochaetes like Marionina spicula and M.achaeta have by far the highest oxygen demands. The outstanding oxygen requirements of meiobenthic oligochaetes have already been stressed by Krüger (1955) for a “very small white enchytraeid” and by Lassèrre & Renaud-

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Mornant (1973) for Marionina spp. The proportionality factor for related to body weight (≈0.8) which Williams et al. (1969) calculated from experiments with freshwater populations of Lumbricillus rivalis was already known from Krüger’s (1955) studies on Enchytraeus albidus, Lumbricillus “lineatus” (probably wrong identification for L.rivalis) and a meiobenthic enchytraeid. Whether it is generally applicable for other (marine) oligochaetes has still to be determined. In (limnetic) tubificids, the physiological background of their often well-developed ability to survive in oligoxic conditions (see p. 240) has been repeatedly studied. Although most of these euryoxybiontic species seem to prefer high oxygen tensions in natural conditions, their metabolism is rarely limited by hypoxia (Alsterberg, 1922; Dausend, 1931; Krüger, 1955; Aston, 1966). For Tubifex tubifex, especially juveniles, high oxygen tensions seem to be even noxious (Fox & Taylor, 1955), as they survived better in 4% than in 100% oxygen-saturated water. In freshwater studies, so many tubificids were found to withstand periods of oxygen deprivation or anaerobiosis that Brinkhurst & Jamieson (1971) generally characterized these forms as facultative anaerobes for restricted periods. Man well (1959) assumes that the body wall in oligochaetes may have a selective potential in maintaining a low internal oxygen pressure even in high external oxygen tensions, a prerequisite for regulated rate of oxygen uptake down to as low external oxygen tension values as 7.5% oxygen saturation (Dales, 1969). Whether the haemoglobin (=erythrocruorin sensu Mangum, 1978) of some marine oligochaete species generally enables these red forms to endure hypoxic conditions better than the often closely related and coexisting species lacking this respiratory pigment, needs detailed scrutinization. Krüger (1955) believes that the presence of haemoglobin in Lumbricillus “lineatus” possibly accounts for its lower oxygen consumption compared with Enchytraeus albidus. For the extremely tolerant species E.albidus and, particularly, Tubifex tubifex, more detailed results on metabolic pathways during anaerobiosis are available. Harnisch (1942) found Enchytraeus to produce “acids” as end-products of probably non-glycolytic processes. Schöttler (1974, 1978), Schöttler & Schroff (1976), and Hoffmann (1978) analysed in detail the metabolic influence of anaerobiosis in Tubifex tubifex. The good adaptation of this species to oxygen deficiencies is demonstrated by the short transition time of 9–12 h for switching from aerobic to anaerobic carbohydrate metabolism which produces mainly propionate and acetate as well as alanine and succinate, but very little lactate. Accordingly, in the naidid Alma emini, during a similar anaerobic degradation of glycogen, lactate could not be detected (Coles, 1970), whereas other oligochaetes accumulated lactate during anaerobiosis (Dales, 1969). That this anaerobic fermentation (interestingly enough very similar to that in endoparasites) is not energetically inferior to the usual aerobic oxidation (Schöttler, 1977) can probably be assumed for many other benthic marine animals which possess corresponding physiological adaptations for life in an intermittently anoxic environment (e.g. Mytilus and Patella, Schöttler, 1974; Pamatmat, 1980), and also for other intertidal annelids like Arenicola (Weber, 1978; Pörtner, Surholt & Grieshaber, 1979). (It should be noticed that in many of the purely physiological papers on “Tubifex tubifex”, “Enchytraeus albidus” or Pachydrilus “lineatus” the basis of species and even

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genus identification, made mostly by non-experts, is more than questionable.)

IMPACT OF DOMESTIC AND INDUSTRIAL POLLUTION DOMESTIC POLLUTION There are few fields in which the gap in our knowledge between marine and limnetic oligochaetes is wider than in pollution studies. In freshwater investigations, the response of “sludge worms” to sewage-induced pollution levels has long been well documented for many areas. Several competent and comprehensive reviews (Brinkhurst & Jamieson, 1971; Aston, 1973b; Brinkhurst & Cook, 1974; Brinkhurst, 1980b; Milbrink, 1980) enable us to restrict ourselves to the few reports dealing with oligochaetes from polluted brackish and marine waters and merely to refer in passing to limnetic studies only in works of general importance or in specialized and less understood fields of pollution. The almost complete absence of references on marine species in the recent reviews on Oligochaeta (in Brinkhurst & Cook, 1980) underlines impressively the lack of data in this field. The rôle which Tubifex tubifex and Limnodrilus hoffmeisteri play in limnetic habitats, namely being the most hardy, ubiquitous inhabitants of heavily polluted waters, still abundant where most fauna are eliminated can apparently, in estuaries, brackish intertidal flats, and subtidal shallow reaches of the sea be ascribed to Peloscolex benedeni, Paranais litoralis and Amphichaeta sannio, and often also to Tubifex costatus. Despite some variations, these species frequently represent an assemblage which in areas of heaviest domestic pollution (often close to sewage outlets) gains almost absolute dominance when all other animals (often with the exception of the polychaetes Capitella capitata and Nereis diversicolor) are exterminated. The generality of this position in pollution gradients is expressed by their ubiquitous presence in a wide salinity range, be it North Sea or Atlantic estuaries (Brinkhurst & Kennedy, 1962; Gray, 1971, 1976; McLusky,Elliott& Warnes, 1978; Birtwell & Arthur, 1980; Diaz, 1980; McLusky et al., 1980), German or Danish tidal flats (Muus, 1967; Otte, 1979), Baltic fjords, bights and inland waters (Knöllner, 1935a; Dahl, 1960; Anger, 1975a,b; Arlt, 1975; Leppäkoski, 1975) or South African polluted shores (Oliff et al., 1967). Despite the generally accepted rôle of this species assemblage in respect to pollution, there is some discrepancy in the ranking particular species. Peloscolex benedeni was found in some investigations only to tolerate “slight pollution” and was absent from the most polluted regions (Leppäkoski, 1975; McLusky et al., 1980), in other studies this species was described as particularly abundant in polluted localities (Knöllner, 1935a; Gray, 1971; Pearson & Rosenberg, 1978) and living close to sewage outlets (Anger, 1975a). Tubifex costatus, living in the Baltic close to its lower salinity thresholds, was absent from the most polluted stations in Arlt’s (1975) study and rarer than Peloscolex benedeni in the very polluted areas examined by Knöllner (1935a). It is, however, found in the more saline Forth Estuary (McLusky, et al., 1980) as the last macrofaunal species in an abiotic area close to a sewer and, in the Thames Estuary, it was particularly aggregated in the most polluted section (Birtwell & Arthur, 1980). Finally, it was considered by Leppäkoski (1975) to occur

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equally often in polluted and unpolluted areas. Less discrepancies seem to exist in the ranking of Paranais litoralis, almost unanimously stated as typical of heavy domestic pollution. Arlt (1975) reports its elevated abundance and dominance with increased pollution, obtaining 14% in unpolluted stations (Amphichaeta sannio: 62.5%), but 63.5% in the most polluted locations (A.sannio: 29%). What are the biological and ecological reasons for the resistance of these particular species to sewage and what favours their increase in abundance when other species disappear? As in their corresponding limnetic relatives, they are particularly adapted to live in and feed on highly enriched organic deposits (Pearson & Rosenberg, 1978) where lack of competition and predation in combination with tolerance for extremely oligoxic, intermittently even anoxic conditions, lead to highly dominating populations of huge densities (e.g. 600,000 Tubifex costatus·m−2 in the Thames Estuary, Birtwell & Arthur, 1980). Hence, to a certain degree the relation of these worms to pollution is anticipated by their extreme respiratory physiology (Brinkhurst & Jamieson, 1971; Aston, 1973a; see p. 240). The ability to endure in grossly polluted sediments may be supported by the beneficial effect which oligochaetes are known to exert by depleting sediments of chemicals by their intense burrowing activity (Chapman, Churchland, Thomson & Michnowsky, 1980; McCall & Fisher, 1980; see p. 245), their irrigation or movements within the tubes (limnetic worms). In this respect, the huge population densities recorded may be highly advantageous in maintaining a considerable outflux of pollutants from the sediment (Aller, 1980). Other tubificid and naidid species from coastal areas have a subordinate rank in pollution studies. The only report on Clitellio arenarius is that of Fraser (1932) who mentions rich populations in the sewage-polluted Mersey Estuary. In oligohaline Finnish coastal areas, Leppäkoski (1975) found, besides Limnodrilus hoffmeisteri, both Potamothrix hammoniensis and Tubificoides heterochaetus in “very polluted” waters ranking these two euryoecious forms as “progressive species of the 1st order”. He found the representatives of this highest pollution category to increase areally and in abundance with the degree of pollution. The abundance of T. heterochaetus in his study compared with its absence in the earlier investigation by Bagge & Ilus (1973) in the same area was interpreted as an indication of severe deterioration of water quality. The latter authors reported Psammoryctides barbatus to occur in the innermost oligohaline areas. In an Indian brackish harbour (annual salinity range 6–33‰), the new tubificid species Monopylephorus indicus and M.waltairensis were recorded to form “patchy mats” of 50×103 to 200×103 ind.·m−2 in areas grossly polluted by sewage and organic garbage (Subba Rao & Venkateswara Rao, 1980). McLusky, et al. (1978) used the occurrence of oligochaetes in relation to other faunal elements when classifying degrees of pollution in the Forth Estuary: in “grossly polluted” areas, oligochaetes in reduced abundance represented the whole fauna; “polluted” regions were abundantly populated by oligochaetes which represented the dominant faunal element among various other groups; and in “largely polluted” areas, density of oligochaetes was reduced compared with other fauna. Based on oligochaetes as a dominant assemblage in polluted waters, both Wharfe (1975) and McLusky et al. (1978, 1980) underlined the value of “community studies” for

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assessing the impact of pollution rather than the establishment of single “indicator species”. Considering the eurytopic nature and often consistent presence of many oligochaete species in polluted as well as in unpolluted waters (see above), the possible impact of pollution-independent factors like substratum (e.g. Anger, 1975b) and seasonal fluctuations on population size, and the poor knowledge we have of life cycles and tolerance ranges in marine species, conclusions based on single species will remain difficult (Brinkhurst, 1966; Aston, 1973b). Consequently, Leppäkoski (1975) evaluated community changes and calculated a “benthic pollution index” on the basis of oligochaete abundance. Howmiller & Scott (1977) in a limnetic oligochaete study had suggested an “environmental index” with three trophic groups which was later modified and related to pollution by Milbrink (1980). According to Howmiller & Scott (1977), Nais elinguis and Ilyodrilus templetoni belong to Group 1 (in slightly enriched, mesotrophic areas), whereas Limnodrilus hoffmeisteri, Tubifex tubifex, and Tubificoides multisetosus in Group 2 characterize highly eutrophic regions, often correlated with massive organic pollution (Milbrink, 1980). In a more general paper, based on data from the tidal flats of the James Estuary (Virginia, U.S.A.), Boesch, Wass & Virnstein (1976) emphasized the importance of “equilibrium species” of high resiliance for the assessment of human disturbances by pollution rather than referring to “opportunistic species”. Subsequent studies of Boesch (1977) and Virnstein (1979) suggest that T.gabriellae may attain a rôle as equilibrium species in the Chesapeake estuaries. Enchytraeids have received even less attention than coastal tubificids and naidids in studies on domestic pollution, although Michaelsen (1926) had noticed that they could survive in sewage-enriched localities only if oxygen is present. Later, Reynoldson (1939a,b, 1947a,b, 1948) in his intensive studies on the fauna of sewage beds stressed the rôle of Lumbricillus rivalis and Enchytraeus albidus in the purification process of waste waster. In his experiments, E.albidus was found to be far more resistant to domestic wastes than Lumbricillus rivalis; this was most noticeable in the restricted tolerance of juveniles. These results contradict Bülow (1957) who stated, without going into more detail, that Enchytraeus albidus was very sensitive to pollution. Williams et al. (1969) also found Lumbricillus rivalis to thrive in filter beds of sewage plants together with Enchytraeus coronatus. Wachs (1963) sampled populations of Lumbricillus rivalis and Paranais litoralis in filter beds of the polluted River Werra (Germany) which is ‘artificially’ saline because of discharges from salt mines (see p. 258). Giere (1970) in his studies in the hygropsammal of a North Sea flat found substantially increasing population densities (3–5 times compared with unaffected stations nearby) in all dominating enchytraeids (L.lineatus, Enchytraeus albidus, Marionina subterranea, M.spicula) near the mouth of a small domestic sewage outlet in a section with high ammonia values of the interstitial water. The high tolerance of Lumbricillus lineatus to sewage contamination in brackish water was confirmed by Arlt (1975) who could not find any significant decrease in populations of this species in a gradient of increasing pollution. Enchytraeus barkudensis from Indian harbour waters also seems to tolerate massive domestic pollution in a brackish environment little oxygenated and heavily exposed to hydrogen sulphide (Subba Rao & Venkateswara Rao, 1980). For other brackish and marine oligochaetes, there are only reports available on the megascolecid Pontodrilus bermudensis from Indian brackish harbour sediments (Subba

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Rao & Ganapati, 1975, Subba Rao & Venkateswara Rao, 1980). This worm thrived in the banks of highly sewage-polluted harbour arms, but decreased in population size in those parts of the harbour affected by ‘chemical’ pollutants (heavy metals, oil, fertilizers). This distributional pattern demonstrates the specificity of oligochaete response to pollution. While domestic sewage often evokes a ‘fertilizing’, positive effect on their populations, chemical effluents can cause severe damage. Hence, terms as vague as “pollution”, repeatedly used by several authors (e.g. Bülow, 1957) are insufficient in ecological considerations of the reaction of oligochaetes to anthropogenic water deterioration. INDUSTRIAL POLLUTION This leads us to the impact of industrial pollution on coastal oligochaetes, a field in its initial phases of investigation. In the older literature there are a few vague statements, again underlining the presence mainly of Limnodrilus hoffmeisteri, a species, apparently highly resistant to a wide variety of agents. This species and Potamothrix hammoniensis resisted the paper and pulp mill wastes in Finnish oligohaline waters (Bagge & Ilus, 1973). Pulp mill wastes in a west of Scotland loch did not impair the development of rich populations of Peloscolex benedeni (Pearson & Stanley, 1979). Lumbricillus lineatus as well as various marine Monopylephorus spp. (Tubificidae) also seem to be highly resistant to pulp mill wastes with low oxygen levels (Coates & Ellis, 1980). These authors even claimed Lumbricillus lineatus as an “indicator species”. Considering the absence of this species in five out of eight effluents studied along the northeastern Pacific coast and its general scarcity in this area, its indicative value is questionable and would seem more generally to reflect stress situations than paper mill pollution in particular. Insecticides and fungicides exerted no detrimental effect on L. hoffmeisteri, whereas polychlorinated phenols proved to be highly toxic both for this species and Tubifex tubifex (Whitley, 1967; Aston, 1973a). In culture experiments (Aston, 1973a) with water heated up to 25°C, egg production of Limnodrilus hoffmeisteri was increased even in low oxygen levels which would explain the higher population abundance (mostly juveniles) downstream of the condenser effluents of a power station in the River Trent (England). Differing results (constant egg production) with Tubifex tubifex make it questionable to generalize about this favourable effect on oligochaete populations, especially with regard to less hardy species. In enchytraeids, Reynoldson (1947b) found a detrimental effect of “acidic” waste waters (pH 3). Their aggravated toxicity at low temperatures was assumed to be based on the longer exposure of maturing stages to the agents. Eggs in cocoons, however, proved to be more resistant, presumably because of the protecting effect of the cocoon wall. Two fields of high environmental risks, heavy metals and oil, have received particular attention in oligochaete pollution studies during recent years, although the impact of metals has. so far, rarely been tested with marine species. Heavy metals Except for some isolated reports (e.g. Brinkhurst, 1962: “metallic poison” in small amounts kills Limnodrilus hoffmeisteri and Tubifex tubifex) most studies on the impact of

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heavy metals are based on Whitley (1967) who found, in detailed tests, that limnetic tubificids were quite resistent to metal ions such as zinc, lead, and nickel, but were highly sensitive to copper. The high toxicity of this metal was also pointed out by Hynes (1974) for tubificids, and substantiated in a recent comprehensive study on Limnodrilus hoffmeisteri and Tubifex tubifex by Chapman et al. (1980). They ascertained not only toxic rates, but focused on uptake and transfer routes of metal including mercury, and the secondary effects on tubificid activity. Uptake occurs mainly by engulfing deposits into the gut, where the metal content was found to be higher than in the tissues. Absorption from the ambient water across the body surface seems of relevance only in high concentrations, since sediment particles contained much higher levels than interstitial or overlying water. Uptake into the animal’s body resulted in a concentration factor of 0.1 to 0.6, varying with the ions tested. Only with mercury, could this factor increase up to 4.0. With regard to this high accumulation, the estimate of Learner & Edwards (1963) that 25% of the mercury in (freshwater) fish is transferred to them from tubificids (and chironomids) must become frightening. Accumulation of noxious metals through tubificids has also been pointed out by Patrick & Loutit (1976). Whereas mercury is known to remain in the food chain for long periods, depuration of other metals like lead from the oligochaete body seems to be rapid. According to Wiese & Jakobi (1975) working on “Tubifex”, after a very short uptake time, lead apparently remains in an “unfixed state” in the animals’ body and is easily re-mobilized for excretion. Stressing the importance of meticulous and calibrated analytical methods in order to obtain reliable toxicity results, Chapman et al. (1980) found in worms preserved in formalin considerably higher metal contents than in the fresh material. They recorded also a substantial impact of heavy metals on respiration rates which was, depending on the agent, in part stimulating and in part inhibitory, without suggesting that this effect was caused by the noxious effects of heavy metals. Whitley (1967) believed that death in tubificids occurred by unnatural ‘precipitation’ of the mucous coating the epidermis which inhibited dermal respiration. A beneficial buffering effect of the mucous layer against toxic matter was also suggested by Reynoldson (1947b) and was particularly effective if the worms were clustered together. The importance of the mucous sheath for the uptake of heavy metals has been impressively quantified in Tubifex tubifex by Fleming & Richards (1982) who could show, using an elegant ‘chelating technique’ with Epoxy-activated Sepharose, that up to 90% of the 65zinc which seemed to be ‘taken up’ by the worms was in fact only surface-associated and only small amounts really ‘internalized’. Other heavy metals (such as cadmium) seem to have a similarly high affinity for the mucus. This burden of mucous-bound agents seems to be only loosely adsorbed so that it can be relatively easily eluted. This could also explain the rapid depuration rates for lead recorded by Wiese & Jacobi (1975, see above). These new results put into question earlier quantitative uptake figures monitored for oligochaetes. On the basis of this new method which, for the first time, allows differentiation between adsorbed and absorbed fractions of noxious heavy metals, many data on uptake quantities should be re-checked. In a detailed physiological study on Enchytraeus albidus, Siebers & Ehlers (1979) analysing the impact of heavy metals on the trans-integumentary absorption of glycine, demonstrated the different action of various metal salts. While exposure to 10 mg·1−1

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iron, lead, zinc, and other ions caused no alteration, and effects of fairly high concentrations of nickel, cobalt etc. remained negligible, 0.1 mg·1−1 mercury and 0.25 mg·1−1 copper, as well as silver and cadmium sharply reduced absorption of glycine. The inhibitory effect in this enchytraeid, identical to that recorded in Nereis diversicolor, is in good agreement with the often recorded toxicity of mercury and copper, and may explain deleterious effects on oligochaetes in estuarine and eulittoral areas, often heavily loaded with industrial wastes. Oil and oil dispersants Coastal areas are increasingly threatened by oil pollution. Continuous discharges from refineries, river water, and accidental oil spills affect the inshore sediments, tidal flats, and beaches all of which are the principal habitats of marine oligochaetes. Extension of our limited knowledge on their response to oil contamination is, therefore, urgently needed. Except for some field studies and one bioassay series (Kasymov & Aliev, 1973) which included observations on oligochaete survival, the only pertinent investigations dealing specifically with marine oligochaetes, both in field research after large oil spills and in laboratory experiments have been conducted by Giere (1979a, 1980a) and Giere & Hauschildt (1979). From field surveys it can be concluded that many of the fresh- and brackish-water oligochaetes known as hardy against pollution are also quite resistant to oil contamination. Peloscolex benedeni was found in British waters near refinery outlets often as the sole surviving member of the macrofauna in high oil concentrations, although population sizes seemed reduced compared with uncontaminated areas, and it was completely exterminated in the immediate vicinity of any discharge (Wharfe, 1975; McLusky et al., 1980; Bagheri & McLusky, in press). From current pollution experiments in a North Sea tidal flat, Rachor (pers. comm.) reports a considerable reduction of P.benedeni populations two months after the oiling of test areas, particularly in muddy sediments with little drainage. This decrease by 70% has to be ascribed mainly to impaired recruitment and reduced development of juveniles. Subsequent tests with an oil-dispersant mixture (Arabian Light: Finasol OSR 2, 10:1) resulted in even more drastic reductions and heavy population fluctuations. From fresh water, McCauley (1966) reported Tubifex sp. as surviving in oily river beds. In a detailed study, Leppäkoski & Lindström (1978) compared faunal assemblages in oligohaline Finnish inland waters near an oil refinery after years of heavy pollution with conditions after a 4-yr post-pollution recovery period. Among the tubificids, Potamothrix hammoniensis was one of the four macrofaunal species having survived the continuous oil stress in the deeper bottoms. In shallow sites, Tubificoides heterochaetus had also maintained considerable population densities which did not increase in dominance through the years of pollution abatement. On the other hand, Tubifex costatus seems to be more sensitive. This truly marine species had completely disappeared from the bottoms near the oil harbour and did not re-colonize the area prior to the 4-yr postpollution period. The available few data on naidids infer relatively good survival for Nais elinguis. In bioassay experiments (Kasymov & Aliev, 1973), specimens from the Black Sea survived in a 300 mg·1−1 water-soluble fraction of Baku oil for not less than 15 days (LT 50) and

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in 3000 mg·1−1 for 9 days. Tentative field data from our own working group confirmed the relatively good survival of this species in polluted tidal freshwater flats of the Elbe Estuary, even when the total population density seemed reduced (Kraft, unpubl. data). Some euryoecious marine enchytraeids like Enchytraeus albidus and Lumbricillus lineatus seem highly resistant both in the experiments of Kaufman (1975) and in the field. According to Bülow (1957) L.lineatus is quite insensitive to “oil residues”. Shortly after the ‘Tsesis’ oil spill, Foberg (1980) could not detect any clearly noxious impact on this species, and in 100 mg·1−1 ‘phenole’-solution Enchytraeus albidus survived for more than 6 days (LT 50, Kaufman, 1975). In more detail, these observations can be corroborated by our own field data on meiobenthic oligochaetes from oil-contaminated beaches. The tubificid Aktedrilus monospermathecus, the enchytraeids Lumbricillus lineatus and Marionina subterranea were found alive after the ‘Urquiola’ spill in considerable numbers in sands with high oil content in which almost all other fauna had perished (Giere, 1979a). Shortly after the ‘Amoco Cadiz’ disaster, we found Phallodrilus prostatus and other, still unidentified oligochaetes even in the beaches directly opposite the wreck thickly covered with the sticky “chocolate mousse’ (Giere et al., in prep.). This rather optimistic picture of oligochaete survival can, however, only be ascertained in studies on the ‘primary impacts’. Re-sampling one year after the spills showed that recovery from a gradual deterioration of oligochaete populations was slow. Reduced diversity and suppressed size of populations indicate longer-lasting impairment than short surveys immediately after the spill may suggest. This long persistence of oil damage was also stressed by Leppäkoski & Kindström (1978). We were interested in the way that these noxious effects would proceed and which biological features would be affected, so we started experiments with Lumbricillus lineatus to determine short-term survival both of adult worms and of developmental stages in various crude oil concentrations, after additions of oil-dispersants and at different temperatures (Giere & Hauschidt, 1979). We also began long-term cultures exposing the worms to sublethal concentrations of ‘Arabian Light’ crude oil (Hauschildt, in press.). L.lineatus seemed especially suitable for these studies due to its wide occurrence, its ease of maintenance and, to its usually massive oil contamination after an accident because of living preferably underneath wrack which gets easily oil-soaked. From experiments with adult worms (Table XIX) one can conclude for short test periods (<14 days) that the species is amazingly tolerant to high oil concentrations. Worms survived even in detritus completely soaked with an oil film. Long-term deterioration becomes particularly obvious after the introduction of an added stress such as high temperature. Even modern oil-dispersants, almost non-toxic in pure solutions at concentrations corresponding to the field situation, are markedly noxious if combined with oil, apparent especially at high (summer)

TABLE XIX Comparison of survival (%) in short-term (14 days) and long-term (60 days) experiments with adult Lumbricillus lineatus: each test was based on 5 times 5, or 2–3 times 20 worms

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10–13°C

26°C

Short-term Long-term Offspring Short-term Long-term Offspring Control

100

100

+

100

100

+

20 000 ppm oil (1:50)

100

100

–

100

0

–

1000 ppm oil (1:1000)

100

100

+

100

100

+

20 000 ppm oil + 2000 ppm Corexit

100

100

–

0 (2 days)

not tested

–

2000 ppm Corexit only

100

100

+

100

100

–

temperatures. Production of offspring is more heavily affected than survival of adults and the increased oil-sensitivity of reproductive stages led to experiments testing the impact on ontogenesis in L.lineatus (Fig. 36). 1000 mg·1−1 crude oil, corresponding to a “medium” contamination (Lindén, 1976), did not recognizably affect developmental stages, sheltered in the cocoons but combination with adequate amounts (1:10) of dispersants (not noxious per se) retarded development by about one third. Higher concentrations of dispersants plus oil, realistically present in contact zones after beach spraying, cause rapid extermination in early developmental stages. In addition, an inverse relation between ontogenetic age and oil sensitivity became apparent. In oil-soaked detritus, in which adults persisted for a long time, embryos initially showed an extremely retarded development, and eventually many of them died in the cocoons. Similar retarding effects in development have been repeatedly reported in stress situations both for this species confronted with extreme salinities (Giere, unpubl. data) and for polychaetes in chemical pollution tests (Bellan, Reish & Foret, 1972). The degree of decreasing productivity and increasing mortality becomes apparent after exposing test populations of L.lineatus to the water-soluble

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Fig. 36.—Impact of ‘Arabian Light’ crude oil and an oil dispersant on embryonic development of Lumbricillus lineatus (18‰ S; 10–13°C): Phases 1–4, in cocoon; Phase ①, sphaerical stage; Phase ②, oval to kidney-shaped stage; Phase ③, oblong to vermiform immobile stage; Phase ④, mobile stage; Phase ⑤, hatching from cocoon; death of animals; data based on 5–10 replicates in each case.

fractions (WSF) of ‘Arabian Light’, which contain particularly large amounts of highly toxic mono- or dicyclic aromates (Table XX, from Hauschildt, unpubl. data). In up to 80 % concentrations of WSF, the cocoon production and viability of offspring are little affected; deleterious effects become clearly visible in full strength solutions. This relatively good resistance is drastically reduced (Table XXI), however, even in adult worms

TABLE XX The impact of water-soluble fractions (WSF) of ‘Arabian Light’ crude oil on productivity, embryonic development and fertility of Lumbricillus lineatus in long-term tests (8 weeks)

% WSF (Arabian Light)

100

80

50

40

20

Control

No. cocoons per 25 worms

94

90

102

110

113

100

% sterile cocoons

17

5

9

8

12

4

% cocoons with non-hatching embryos

11

1.2

1.3

2.6

3.3

3.1

% hatched juveniles

67

86

83

80

78

86

TABLE XXI The impact of water-soluble fractions (WSF) of ‘Arabian Light’ crude oil in combination with a dispersant on mortality and productivity of Lumbricillus lineatus. Comparison of short- and long-term results

% WSF (oil+Corexit, 10:1)

100

75

50

25

Control

Mortality (%) adult worms within 1 week

36

16

0

0

0

within 6 weeks

76

52

36

0

0

9

8

15

86

≈100

No. of cocoons per 25 worms

and in their cocoon production, when oil-dispersants are added (data tor Corexit 7664, Esso, are largely identical to those for Finasol OSR 2, Fina).

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When the animals were maintained in oiled Fucus, they fed intensely on this debris and so incorporated large quantities of oil into their bodies. Scrutinizing this uptake, we found oil not only passing through the gut, but being transferred into the chloragocytic layer around the intestine, which clearly became blackened. Microscopic examination showed large numbers of tiny dark globules in the chloragogenic cells, normally a tissue for lipid metabolism, storage, and excretion of metabolic wastes (Fig. 37). Fluorescence optical spectra first proved these substances to have been chemically transformed by metabolic processes (Hauschildt, pers. comm.); their later fate is as yet unknown. Uptake of mineral oil in relation to lipid metabolism is also known from experiments with other animals (Rossi & Anderson, 1977). Current research on long-term effects (so far six generations) in Lumbricillus lineatus cultures exposed to from 20–80% WSF (Hauschildt, in press) can be briefly summarized to result in depressed growth with massive retardation of generation time in the F1, but with return to normal spans in the following generations. While overall cocoon production seemed to be

Fig. 37.—Incorporated oil globules (black dots) in the chloragogenic tissue of Lumbricillus lineatus: b. s., blood sinuses; chl., chloragogenic tissue; coe., coelomic cavity; d. mu., dermal musculature; int., intestinal

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wall; fixation, osmium-tetroxide and stain, Toluidine-Blue; (from Giere & Hauschildt, 1978).

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only little impaired by the hydrocarbons, fertility of eggs, formation of normal embryos and hatching rates decreased markedly through the generations. In combinations of WSF’s with the dispersants ‘Corexit 7664’ and ‘Finasol OSR 5’ (10:1), detrimental consequences for fertility and normal embryonic development were far more expressed. The highly toxic effects even of these ‘modern’ agents of ‘little toxicity’ were scarcely evident in the ‘starting generation’, but became revealed particularly after the second and third generation, so that continuous culturing was only possible in the weakest solutions (20% WSF). These results impressively demonstrate the need for long term studies and the severe limitation of the frequently obtained ‘immediate impact data’. In Hauschildt’s experiments, a remarkable diversity of results was found in some cohorts, reminiscent of the flexibility of this species in normal development (see p. 191), in salinity tolerance tests (see p. 260), and in the inconsistent survival of adults in oil-soaked debris in which some worms were always killed while others persisted (see above). In other experiments interstitial marine oligochaetes were tested for the first time. Adult Marionina subterranea, regularly found to be fairly unaffected in field surveys after spills, tolerated 1000 mg·1−1 oil without significant losses during 14-day tests. In chemically dispersed oil mixtures (Corexit 7664, Finasol OSR 2, OSR 5) survival decreased to ≈70–90%. Despite this impairment of survival rate, littoral oligochaetes appeared in short-term tests to be amazingly tolerant to oil and (in low concentrations) even to dispersants. This corresponds well with many of our field observations immediately after oil spills. After accidents, beaches and high-water lines are, however, usually exposed to thick oil layers and, regrettably, also to intense contact with dispersants in order to ‘clean-up’ the beach sites. Under these conditions, even the hardy oligochaetes will be largely exterminated. Moreover, our long-term studies through several generations clearly show the slow and often barely susceptible secondary effects which result in a gradual decrease and imbalance of populations, particularly in the unstable upper eulittoral with its array of additional stress factors (e.g. high temperature). This again demonstrates the inadequacy of rash mortality assessments soon after the spill, and emphasizes the need for more longterm studies (Bellan, et al., 1972; Leppäkoski & Lindström, 1978). Particularly in muddy flats, one main habitat of marine oligochaetes, persistence and threat of oil in the sediment lies in the range of decennials (Chassé, 1978; Burns & Teal, 1979; Gunkel & Gassmann, 1980) exerting a continuous oil stress by sublethal pollution. Details of the negative action of oil or dispersants on the oligochaetes are little understood. Apart from toxic effects on the worm’s physiology and decreasing oxygen content of the bottom, the mechanical impact on sediment structure probably contributes to population decreases. Near refinery outlets, bitumen-like strata prevented tubificids

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from burrowing (Leppäkoski & Lindström, 1978). Wharfe (1975) reported Peloscolex benedeni to burrow in oil-polluted muds but to remain much nearer the surface than is usual and supposed that the presence of oil might hinder the worms’ burrowing activity. Mesopsammic oligochaetes like Marionina subterranea living in relatively exposed and permeable beaches, where oil is usually soon flushed away by the waves, are more jeopardized by strong thigmotaxis, which enhances the risk of animals getting smeared by oil droplets absorbed on grains and in crevices. Direct contact of the epidermis and oil was highly toxic in all cases (Giere, 1979a). The striking ability of Lumbricillus lineatus to triturate oiled wrack and to incorporate and biochemically transform oil, as shown above, may be an important contribution to the natural decomposition and removal of oil after spills considering the masses of algal debris consumed by the rich worm populations (Giere & Hauschildt, 1979). Moreover, enhancement of surfaces by browsing off the detritus and crushing down oil slicks adsorbed to algal mats into smaller hydrocarbon particles by the worms’ activity will further intensify bacterial oil degradation (Tenore, 1977; Gardner, Lee, Tenore & Smith, 1979; see p. 207) in the shoreline. In a survey of the response of oligochaetes to oil, the strange observations of Spies & Davis (1979) and Davis & Spies (1980) should not be omitted. They investigated the benthic fauna of natural oil seeps off the Californian coast and found marine sublittoral oligochaetes (the tubificids Limnodriloides verrucosus and L.monothecus) to be particularly frequent in heavily contaminated sandy seepage areas (3 mg oil·g−1 sediment). In some chemical aspects, the n-alkanes of the seep oil, however, were evidently microbially degraded and, thus, perhaps detoxified. The rich bacterial biomass plus the possible adaptation of the fauna in seepage areas through the ages may explain not only the absence of any noxious impact of this natural oil spill but also the support of rich worm populations feeding on deposits (14 ind.·core−1 at a seepage station compared with ≈1 at an uncontaminated site).

THE RÔLE OF OLIGOCHAETES IN MARINE ECOSYSTEMS; COMPILATIONS, CONCLUSIONS, AND SPECULATIONS Considerations of the mode and intensity of interactions of oligochaetes with their marine environment have to be differentiated in relation to their wide zonation and affiliation to various biocoenoses. With regard to the marine biota as a whole, oligochaetes are of subordinate importance both numerically and ecologically. This is certainly true for the sublittoral meiobenthic tubificids in spite of their intensively proceeding radiation. Although there exists highly interesting anatomical, reproductive (e.g. external attachment of spermatophores; Erséus, 1980b), and nutritive patterns (uptake of dissolved organics, see p. 211), and although their ecology is still widely unknown, one cannot ascribe to them a major rank regarding their low population densities and high competition with other benthos in this zone. Approaching shallow non-tidal beaches or eulittoral and estuarine environments, marine oligochaetes increase both in abundance and ecological relevance. They seem better adapted than most thalassogenic forms to this zone of heavy physiographical

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fluctuations, often unpredictable and severe. Characteristically, their diversity decreases in these extreme biotopes while, particularly in naidids and the larger macrobenthic tubificids, their abundance in favourable times can exceed that of all other groups. This growing dominance in coastal areas is certainly related to the predominantly brackish conditions which exclude most marine competitors (Diaz, 1980), whereas oligochaetes, of limnetic origin, thrive in these regions of high natural and often also anthropogenic organic enrichment. Their burrowing and feeding activity has a considerable impact in structure, nutritive value, chemistry, and re-mineralization (nutrient recycling) of the sediment. Oligochaete bioturbation, known to deplete considerably the pollutants from sediments, attains a particular importance because especially in areas stressed by wastes, masses of oligochaetes often represent the sole surviving fauna. Thus, in polluted coastal areas, oligochaetes, according to McLusky et al. (1980) may be “energetically at least as important as the (remaining) macrofauna” (see also Leppäkoski, 1975). Proceeding to the high-water line and supralittoral beaches, two oligochaete assemblages, largely different from the ones described above, will be encountered. In the mesopsammon of moist sands, interstitial enchytraeids and tubificids, often provided with coelomic pores and glands excreting adhesive substances (Jansson, 1961; Giere, 1971c), can represent a major meiobenthic group (Fenchel, Jansson & Thun, 1967; Giere, 1975; Munro, Wells & McIntyre, 1978), particularly in the upper layers and the top shore of detritus-rich beaches. In and beneath wrack beds, which often accumulate along these shores, macrobenthic enchytraeids represent one of the few marine elements in a transitional marine-terrestrial community. Through their huge population densities these well-adapted oligochaetes attain a considerable importance for re-working and triturating plant material washed ashore, thus intensifying also microbial activity (Giere, 1975; Tenore, Tietjen & Lee, 1977). Which factors enable oligochaetes to successfully maintain themselves in these ecologically rigorous biotopes? Despite their frequently worldwide occurrence and euryoecious nature, our scant knowledge about their autecology and population dynamics renders definite answers quite difficult. As evidenced from tolerance-preference experiments, their euryoecious nature provides most coastal oligochaetes with a capacity to endure, in temperate regions, even extremes of physiographical fluctuations. This hardy and eurytopic nature does not blur specific tolerance differences for various factors and particular adaptations, e.g. to low oxygen levels in many tubificids, to low water content in many enchytraeids or to restricted dietary demands. A differentiated discriminatory potential, a variety of biotic interrelations, a diversified food range, and a well discernible preference behaviour, flexible enough for long-term adaptations to new ecological situations, all suggest that in these worms control mechanisms are beyond the simple pattern of capacity reactions. In combination with often strong population fluctuations due to seasonal shifts and reproductive periods, these features will contribute to the well-known patchy distribution pattern and changing abundance of oligochaetes (Nielsen, 1954). Nutritional aspects lead to the question in which way oligochaetes in general are incorporated in the marine ecosystem and, more specifically, which relation exists between meiobenthic worms and macrobenthic predators (Fig. 38). Scrutinizing the food demands even of the so-called ‘omnivorous’ deposit-feeders, it becomes clear that really

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digestible food is by far not as abundant as the rich detritus supply makes one believe. Thus, nutritive preferences and trophic specializations can determine oligochaete

Fig. 38.—Marine oligochaete zonation related to their differentiated incorporation into the food spectrum of carnivores.

distribution pattern even in an environment of apparently abundant food supply (Backlund, 1945; Giere, 1975). Accessibility is one of the key words for the oligochaetes’ rôle in the food web. Whereas predation on the intertidal tubificids is reduced by their endobenthic life (Diaz, 1980), naidids with their more epibenthic activity are frequently subject to severe exploitation. Finally, enchytraeids, occurring mainly in the supralittoral, escape predation by aquatic forms and, often sheltered by thick wrack layers, suffer even relatively little from sea birds. This relation between zonation of oligochaetes and their utilization as food is illustrated in Figure 38. Despite the variety of possible predators (see p. 213), we are inclined to the view that the bulk of oligochaete biomass will be recycled via organic decomposition (Giere, 1975). In estuarine habitats, Diaz (1980) recognizes a more important participation of oligochaetes in the food transfer to higher carnivorous levels, suggesting a close connection between polychaete dominance in marine and oligochaete prevalence in limnetic biotas. As with meiobenthos in general, the rôle meiofaunal oligochaetes play as a food source for macrofaunal consumers, thus representing links which mediate between microorganisms and higher predators, is not unanimously recognized. Observations on intense interactions between meiofaunal members, viz. interstitial oligochaetes and predators like turbellarians (Dörjes, 1968; Radziejewska & Stankowska-Radziun, 1979), but less

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meiofaunal consumption by macrobenthic grazers, follow the ‘closed-circuit’ line of reasoning suggested mainly by McIntyre (1969, 1971), and later experimentally confirmed by McLachlan (1977a), Tenore (1977), and Reise (1979) for particular ecosystems. Other authors report from various habitats, and also on the basis of experiments, that oligochaetes as well as other vermiform meiofauna were subject to massive predation, thus allowing the generaliz-ation that meiofauna is effectively controlled by macrofauna predation and is firmly integrated into the food web leading to macrofauna (Elmgren, 1976; Bell & Coull, 1978; Gerlach, 1978). Coull & Bell (1979), reviewing these contradictory arguments, concluded that in sandy bottoms meiofaunamacrofauna interactions are negligible, whereas in muddy substrata with their more horizontal colonization structure, grazing pressure by macrofauna is high. Hence, the rôle of oligochaetes in the marine food web seems to depend largely on local ambient conditions and on the species assemblage involved. This implies that results will vary considerably and much more knowledge on the biotic interrelations of marine oligochaetes is required before general conclusions, exceeding the level of local studies, can be drawn. In spite of the high physiological stress acting in temperate shallow bottoms and shores upon the inhabitants, there is in our opinion not enough conclusive evidence that many oligochaetes, well adapted to these rigid conditions, are indeed predominantly physically controlled (Giere & Pfannkuche, 1978) as usually assumed for residents of ‘extreme’ biotopes. Referring to other macro- and meiofauna, a massive impact of biological interferences is confirmed for tidal flats by recent studies of Reise (1977, 1978), for atidal beaches by Hulings & Gray (1976), for shallow non-tidal bottoms by Arntz (1980), and for polluted estuaries by Gray (1976). That biological control mechanisms are also substantially operative in stressed areas is only in part a contradiction of Sanders’ (1958) stability-time hypothesis. This becomes clear from a recent summary paper of Sanders (1979) in which the author emphasizes the severe competition and predation due to imbalanced biological interaction in regions of high physiological stress, a corollary often overlooked in discussions on this hypothesis. In subtropical and tropical shores, extremes of abiotic factors, viz. salinity and temperature, and occasionally also pH, can approach the upper tolerance limits and, notably in a combination of stresses, directly define the tolerance capacity of oligochaetes, thus creating a predominantly physically controlled system. We have to exempt from our suggestions environments dominated by chemical pollution, a factor to which limits of adaptation are narrower and which, in severe cases, becomes controlling even for the largely tolerant oligochaete species (Brinkhurst & Kennedy, 1962; Elmgren, 1976; Hulings & Gray, 1976). The difficulties of characterizing population dynamics of marine oligochaetes in terms of Sanders’ (1958) control principles are paralleled by similar problems to ascribe the ecological structure of the worms to categories typically associated with r- and kstrategists. Extreme width in their tolerance capacity, flexibility in embryonic and juvenile development, seasonal breeding periods, widespread occurrence of relatively few species, and high genetic variability (Lumbricillus) are apparent opportunistic adjustments of r-strategists to a physically controlled environment. Relatively low egg number, brood protection in cocoons with absence of free larvae are, however, features

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apparently better suited to k-strategists, but should probably be interpreted in relation to the limnogenic nature of marine oligochaetes. Moreover, brooding can be taken as a favourable adaptation to life in physically rigid and often polluted environments (Gray, 1976) warranting high survival in organisms which are particularly valuable for exploiting new, highly disturbed environments abandoned by more sensitive fauna (Pearson & Rosenberg, 1978). In sublittoral meiobenthic oligochaetes, the trend towards k-strategy is far more developed than in their common eulittoral relatives. Restriction to particular habitats (e.g. to coralline sands, Amphioxus sands), intense speciation with many endemists (e.g. in Phallodrilinae), and breeding periods not adjusted to the seasons (Erséus, 1976a) have to be considered as evolutionary steps leading to more restricted, stenoecious conditions and to a truly marine life remote from their limnetic ancestors. For a better ecological and biological evaluation of these ‘new’ marine invaders, further research is needed preferably on their autecology (life cycles, production, response to natural stress, and anthropogenous disturbance), on their physiology, and on their linkage to other members of the marine ecosystem (food relations). Only then will consideration on their rôle in the marine biota loose its speculative character.

REFERENCES Abrahamsen, G., 1969. Oikos, 20, 54–66. Abrahamsen, G., 1973. Pedobiologia, 13, 6–15. Abrahamsen, G. & Strand, L., 1970. Oikos, 21, 276–284. Åkesson, B., 1975. Pubbl. Staz. zool. Napoli, Suppl. 39, 377–398. Albert, R., 1975. Mitt. hamb. zool. Mus. Inst., 72, 79–90. Alderdice, D.F., 1972. In, Marine Ecology, Vol. I, Pt. 3, edited by O.Kinne, Wiley Interscience, London, New York, 1659–1722. Allen, R.R., 1950. N.Z. Sci. Rev., 8, 89, only. Aller, R.C., 1980. In, Marine Benthic Dynamics, edited by K.R.Tenore & B.C. Coull, University of South Carolina Press, Columbia, 285–308. Alongi, D.M. & Tietjen, J.H., 1980. In, Marine Benthic Dynamics, edited by K.R. Tenore & B.C.Coull, University of South Carolina Press, Columbia, 151–166. Alsterberg, G., 1922. Acta Univ. lund., N.F., Avd. 2, 18, 1–176. Andrássy, I., 1956. Acta zool. hung., 2, 1–15. Anger, K., 1975a. Merentutkimuslait. Julk./HavsforskInst. Skr., 239, 116–122. Anger, K., 1975b. Helgoländer wiss. Meeresunters., 27, 408–138. Ankar, S., 1977. Contr. Askö Lab., 19, 62 pp. Ankar, S. & Elmgren, R., 1975. Merentutkimuslait. Julk./HavsforskInst. Skr., 239, 257– 264. Ankar, S. & Elmgren, R., 1976. Contr. Askö Lab., 11, 115 pp. Appleby, A.G. & Brinkhurst, R.O., 1970. J. Fish. Res. Bd Can., 27, 1971–1982. Arlt, G., 1973. Wiss. Z. Univ. Rostock Reihe Mathematik und Naturwissenschaften, 22, 1141–1145. Arlt, G., 1975. Merentutkimuslait. Julk./HavsforskInst. Skr., 239, 272–279. Arndt, E.A., 1973. Oikos, Suppl. 15, 239–245. Arndt, E.A. & Nehls, H.W., 1964. Z. Fisch., N.F., 12, 45–73.

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THE BIOLOGY OF SANDY-BEACH WHELKS OF THE GENUS BULLIA (NASSARIIDAE) A.C.BROWN Department of Zoology, University of Cape Town, Rondebosch, South Africa

Oceanogr. Mar. Biol. Ann. Rev., 1982, 20, 309–361 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION The stenoglossan family Nassariidae belongs to the most advanced superfamily of neogastropods, the Buccinacea. It comprises a number of small to medium-sized whelks, living almost entirely on soft substrata and scavenging animal matter. The foot is characteristically very broad and thin, and often ends in two small posterior cirri. The distribution of the family is world-wide. Nassarius and its close relatives tend to be confined to mud, although a few species shelter among loose rocks (Kilburn & Rippey, 1982), while species belonging to the Bullia group (Bullia, Dorsanum, Buccinanops) are normally found on sandy substrata, particularly in sheltered bays. Some members of the genus Bullia are, however, adapted to life in the surf zones of high-energy sandy beaches and it is with these that the present review is chiefly concerned. These sandy-beach whelks are most common in, although not entirely restricted to, the Southern Hemisphere and occur in temperate, subtropical and tropical waters. Their taxonomy, and particularly the status of the alleged genera and subgenera, still leaves much to be desired and the distribution of most species is poorly known. New species continue to be described (Kilburn, 1978) and the taxonomy of known species to be revised (Cernohorsky, 1977; Brown, 1979b). The genus Bullia was erected by Gray (in Griffith, 1833) for the purpose of distinguishing a group of species which he thought to be intermediate between Buccinum and Terebra and characterized by a remarkably wide expansion of the foot. From the chaos of prosobranch systematics which prevailed during the latter half of the nineteenth century emerged the family Nassidae (now Nassariidae), containing the genera Bullia and Dorsanum, the former being distinct from the latter in having no eyes. Large numbers of Bullia species were described, largely from southern Africa, empty shells, often beachworn, being accepted as sufficient evidence for the erection of a new type. Shell pattern was held to be specific and colour-varieties were legion. The resulting confusion has only been resolved for the South African region by Barnard (1959) and Kilburn & Rippey (1982), although some workers continue to regard Dorsanum as a subgenus of Bullia,

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following the opinions of Nicklès (1950), Buchanan (1954), Gauld & Buchanan (1956), and others. Even Brown, Ansell & Trevallion (1978) attributed B. melanoides to the subgenus Dorsanum; a double error as Bullia melanoides is a true Bullia and does not belong to the Dorsanum group of species. In addition to the recent species, a number of fossil forms are recorded (Wenz, 1938; Barnard, 1962). To attribute these shells to any genus within the Nassariidae would, however, appear to be the merest guesswork. Useful descriptive and experimental work on the Bullia group of whelks has been carried out in South America, India and South Africa, notably at the Universities of Cape Town and Port Elizabeth. Crichton (1942a,b) reported briefly on the habits and lifehistories of three Bullia species from the east coast of India, while Ansell & Trevallion (1969, 1970) and Ansell et al. (1972) have described the behaviour and ecology of B. melanoides from southwestern India. Penchaszadeh (1971a,b, 1973) reported on the reproductive habits and aspects of the ecology of Dorsanum and Buccinanops from Argentina. Brown (1961a, 1964d, 1971a,b) dealt with the behaviour and ecology of Bullia laevissima, B. rhodostoma, and B.digitalis on South Africa’s Cape Peninsula beaches and the ecology of the latter two species, as well as that of B. pura, has received attention on Eastern Cape beaches from staff and students of the University of Port Elizabeth (McLachlan, 1977a,b; McLachlan, Cooper & Van der Horst, 1979a; McLachlan & Van der Horst, 1979; McLachlan, Wooldridge, Schram & Kühn, 1979b; McLachlan, Wooldridge & Van der Horst, 1979c; Ansell & McLachlan, 1980; McGwynne, 1980; Dye & McGwynne, 1980). The ecology of B. digitalis on South Africa’s west coast has been studied by Bally (1981). Recent work at the University of Cape Town has concentrated on the population of B. digitalis to be found at Ou Skip, on the west coast of South Africa just north of Table Bay. Among other investigations, the oxygen consumption of these whelks has been studied under a variety of controlled conditions (Brown et al., 1978; Brown, 1979a,b,c,d; Brown & Da Silva, 1979; Brown & Meredith, 1981). This work has led to the construction of an activity (time-energy) budget and an estimate of the cost of free existence of adult animals (Brown, 1981, 1982). Locomotion and pedal extension and retraction have been studied by Brown (1964b) and by Trueman & Brown (1975, 1976), while aspects of pathology and the rôle of the haemocytes were considered by Brown & Brown (1965). The effects of pollution on Bullia have received attention from Brown (1964c), Brown & Currie (1973), Brown, Baissac & Leon (1974), Brown, Davies & Young (1982), Cuthbert, Brown & Orren (1976a,b) and Golombick & Brown (1980). The structure and function of the osphradium have been described by Brown & Noble (1960) and Newell & Brown (1977), while Meredith & Brown (in press) and Meredith (in prep.) have been concerned with a scanning electron micrographic study of Bullia radulae and the ultrastructure of the pedal musculature. There are also a number of unpublished theses and student projects dealing with Bullia, as well as short published notes. In addition there are valuable observations in papers dealing mainly with other matters. A full bibliography of the genus Bullia, including unpublished work, has in fact been prepared (Brown, 1979e). The results of much of the above work invite comparison with those gained from mud snails of the closely-related genus Nassarius, on one hand, and with sandy-beach naticids, such as Polinices, on the other. The present review attempts to bring together all

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the published, and most of the

Fig. 1.—Bullia digitalis (photograph by P.F.Newell, previously unpubl.).

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Fig. 2.—Photographs of living Bullia melanoides illustrating the degree of expansion and flexibility of the foot when extended to act as an underwater sail in surfing (A-E): the characteristic posterior cirri and the small operculum may be seen; F, the posture during burrowing, with the siphon extended vertically in front; (from Ansell & Trevallion, 1969).

unpublished, work on Bullia, other than the systematic and zoogeographical literature,

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and essays some comparison with other molluscs, where appropriate.

GENERAL ECOLOGY AND BEHAVIOUR THE SURFING HABIT AND RESPONSES TO CURRENTS From the points of view of ecology and behaviour, whelks of the genus Bullia may conveniently be divided into two groups—those which are predominantly intertidal and exploit wave action and currents by spreading their broad, thin, agile feet and surfing up and down the shore, and those which are chiefly or entirely subtidal and do not surf. Many localities around the South African coast have both surfing and non-surfing species and some bays harbour several species of each category. Thus in False Bay and in Algoa Bay B. digitalis, B. rhodostoma, and B. pura all occur intertidally and exploit the surf, while offshore may be found the non-surfing B. laevissima, B. tennis, B. annulata, B. diluta, and B. callosa (Brown, 1971b; McGwynne, 1980). Brown (1961a) made a comparative study of two of these species, the intertidal B. digitalis, a common and conspicuous surfing species on sandy shores of the west and south coasts of South Africa, and B. laevissima, a shallow-water, non-surfing whelk occurring along roughly the same length of coast. The latter species is found intertidally only on low-energy beaches not frequented by B. digitalis. A crucial difference with regard to response to water currents was shown to exist between the two species. The surf-loving B. digitalis (Fig. 1) emerges actively from the sand in response to currents, turns on its back and spreads its foot, turning and twisting it from side to side, a behaviour pattern which encourages transport by waves and wash. In a directional current the whelk may turn onto its side, with the apex of the shell pointing into the current, and move actively with the current by pulling the shell over the sand with the margin of the foot. In contrast, B. laevissima responds to currents by burrowing deeper into the sand, emerging only when the currents cease or are drastically reduced. Field experiments confirmed the importance of these contrasting responses in the different habitats and lifestyles of the two species, responses which draw B. digitalis towards high-energy intertidal sands, while inhibiting B. laevissima from approaching the shore except where wave action is minimal. Emergence from the sand and surfing in response to currents has been noted in other intertidal Bullia species, including B. vittata, B. melanoides (Fig. 2), B. rhodostoma, B. pura, and B. natalensis (Crichton, 1942b; Ansell & behaviour resembles that of the toxiglossan whelk Terebra salleana as Trevallion, 1969; Brown, 1971b, unpubl.; McLachlan et al., 1979c). The described by Kornicker (1961). Another important difference between the two species studied by Brown (1961a), relevant to their different responses to currents, is that the subtidal Bullia laevissima has a relatively massive and heavy shell, and consequently a higher specific gravity, than has B. digitalis. The latter is thus more suited to surfing. In this regard, it is interesting that B. rhodostoma, which surfs higher up the shore than B. digitalis, has a still lighter shell and lower specific gravity than B. digitalis (McGwynne, 1980).

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TIDAL MIGRATIONS Tidally regulated migrations up and down the intertidal slope appear to be the rule rather than the exception for the aquatic macrofauna of high-energy sandy beaches. Bivalves such as Donax, cirolanid isopods, Emerita, the prosobranch Terebra, and a variety of other forms all display such behaviour (Newell, 1979). The surfing species of Bullia are no exception and tidal migrations have been noted by Crichton (1942b), Brown (1961a, 1971b), Ansell & Trevallion (1969) and Trevallion, Ansell, Sivadas & Narayanan (1970). McLachlan et al. (1979e) have studied this phenomenon quantitatively on the rich Maitland River Beach, in the Eastern Cape Province of South Africa (see Fig. 3). They found that B. rhodostoma shows a

Fig. 3.—Tidal migrations of Bullia rhodostoma over 24 h on Maitland River Beach: numbers at each sampling time are plotted on a percentage scale; thick line on the time scale indicates the period of darkness, while the graph (●) shows the upper limit of wash at each time; (from McLachlan et al., 1979c).

pronounced tidal cycle of migration both during the day and at night, the centre of gravity of the population moving up and down the shore a distance of some 45 m during spring tides. Large individuals migrate a greater distance than small ones, as at low tide a sizebased zonation is apparent, with the smallest whelks occupying the uppermost zone. Such zonation is not in evidence at mid- or high tide. B. digitalis, on the same beach, was also shown to migrate up and down the shore, again covering distances of about 45 m, but keeping lower down the beach than B. rhodostoma, so that it was never exposed to air. These observations do not necessarily apply to all populations and indeed Brown

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(1961a, 1971b) has found that on the Cape Peninsula beaches he studied, although tidal migrations occur, the populations lag behind the tide as it rises and avoid the upper part of the intertidal zone, the animals burying themselves in the sand during high water. These differences in behaviour may be partly explained by differences in beach profile, for the beaches studied on the Cape Peninsula were all typically saucer-shaped in profile, the slope increasing up the shore (Brown, 1971a), while Maitland River Beach studied in the Eastern Cape “shows a very uniform slope from the waters edge to the base of the dunes” (McLachlan et al., 1979c). Such differences in profile may well influence the extent of tidal migrations, particularly as there appears to be no intrinsic migratory rhythm in Bullia (Brown, 1961a; McLachlan et al., 1979c), so that these movements must result entirely from the animal’s behavioural responses to changing conditions as the tide moves up and down the intertidal slope. In this regard it is significant that beaches with an overall steep profile do not harbour Bullia intertidally, although subtidal populations of surfing species may occur and invade the intertidal sands after the profile has been flattened during storms (Brown, 1971a). This is especially true of B.digitalis on South Africa’s west coast. Brown (1981a) has also observed that the length of the “activity period” in B. digitalis varies considerably from beach to beach, “probably depending on beach slope, wave action and other factors”. McLachlan (1980) has found that, within the range of wave action on the beaches studied by him, B. rhodostoma tends to move upshore on relatively exposed beaches, to occupy a higher position than on more sheltered beaches. This tendency depends on the strength of the wash and is offset by increasing beach slope. On Cape Peninsula and South African west coast beaches, the positions occupied by the various Bullia species differ from day to day with changing conditions of wave action (Brown, unpubl.). The ability of B. rhodostoma to migrate higher up the beach than B. digitalis or B. pura is undoubtedly due to greater surfing efficiency. McGwynne (1980) has demonstrated that B. rhodostoma surfs at a faster rate than the other two species, averaging about threequarters of the flow rate of the water under experimental conditions. B. digitalis and B. pura average a little less than half the flow rate. The rate of surfing between large and small whelks of the same species is not, however, significantly different. DISTRIBUTION ON THE BEACH As noted above, the Bullia populations on Maitland River Beach and other beaches in the vicinity of Algoa Bay show a marked zonation, B. rhodostoma occupying a zone above that of B. digitalis and B. pura, while the juveniles tend to be zoned higher than the adults at low tide, at least on some beaches. A zonation of sorts also occurs on Cape Peninsula beaches but is far less marked. On Muizenberg Beach, False Bay, B. rhodostoma and B. digitalis tend to be separated by a long-shore segregation (Brown, 1971b). The animals occur in well-defined groups, separated by stretches of beach in which no whelks are to be found. Each group normally consists of a single species and is often confined to a narrow size range. As the whelks do not appear to be gregarious, as are some species of Nassarius (Crisp, 1969), the most likely explanation for this distribution is a sorting by the waves and currents during surfing (Brown, 1971b). On some other beaches, Bullia populations are largely confined to limited stretches, and this too can be related to water

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currents. B. rhodostoma is generally confined to the intertidal slope but B. digitalis, although a surfing species, may occur in some numbers offshore. Christie (1976) found this species at a depth of 5 m and more in Lambert’s Bay and according to Brown (1961a) its maximum recorded depth is 20m. Under certain circumstances B. digitalis may, however, be found in still deeper water, particularly off exposed beaches with steep slopes, where it can at best only occur intermittently (Brown, unpubl.; J.C.Allen, pers. comm.). McGwynne (1980) reports that the B. digitalis population tends to move offshore from Eastern Cape beaches during winter. The distribution of Bullia does not appear to be influenced by sedimentary characteristics, differing in this respect from several other members of the macrofauna inhabiting the same beaches (Brown & Talbot, 1972; Brown, 1973). Bally (1981) has found no correlation between the distribution of B. digitalis and grade of sand on South Africa’s west coast, and Brown (1961a) showed in the laboratory that particular grades are not preferred over others. The surfing species of Bullia occur only within a certain range of exposure to wave action although, where this is heavy, it is not possible to distinguish between wave action and beach slope as the chief factor excluding them (Brown, 1971a,b). Where wave action is slight they are commonly replaced by non-surfing species such as B.laevissima (Brown, 1961a). Bullia is also absent from many South African beaches which fall well within the range of wave action and beach slope required. Such beaches are often fringed or ringed with kelp such as Ecklonia and at times may be subjected to large deposits of kelp material (Brown & Talbot, 1972). CONTACT BEHAVIOUR Ansell & Trevallion (1969) observed a behavioural response in Bullia melanoides which has not been noted in other species of the genus; when two crawling individuals come into contact, one reacts by turning suddenly onto its side, remaining inverted for some time. In the field, this will encourage transport in the surf, thus tending to maintain dispersion of the population. The response is similar to that exhibited by several other Nassariidae and some members of the Buccinidae in response to predators (Feder, 1967). FOOD AND FEEDING The Nassariidae are typically carnivorous scavengers, although there are a number of partial exceptions. One North American Nassarius appears to prefer the eggs of annelid worms to dead animal matter, while another supplements its diet by scooping up mud and digesting the micro-organisms it contains (Kilburn & Rippey, 1982). Curtis & Hurd (1979) have shown that Nassarius obsoletus is an obligate omnivore, demanding both plant and animal material in its diet if growth and reproduction are to be maintained, while Palmer (1980) has found that N. kraussianus will graze on algae or Zostera and will survive for at least two months on these plants without access to carrion. The Bullia group, including Dorsanum and Buccinanops, in general conform to the nassarid pattern of feeding mainly on carrion. Bullia will, however, accept a wide variety

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of animal matter, including Cnidaria, annelids, other molluscs, crustaceans, insects, tunicates, fish of all kinds,

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Fig. 4.—Bullia rhodostoma feeding on a stranded coelenterate medusa: the whelk near the centre of the picture has its proboscis thurst deeply into the carrion; the remarkable length of this organ when fully extended may be seen; (from Brown, 1971b).

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birds and mammals (Brown, 1961a, 1964d, 1971b; McLachlan et al., 1981), although the latter two groups are not favoured. On the Indian Ocean shores of South Africa, stranded Cnidaria, particularly siphonophores such as Physalia, and a variety of jellyfish (Fig. 4) form the staple diet of the intertidal Bullia species. In the case of Physalia, the float is avoided and the tentacles are attacked first, despite the nematoblasts they contain. Living, freshly-stranded Physalia seem to be preferred to those which have been lying for some time in the sun. In the case of jellyfish, the bell is usually attacked, the whelk’s proboscis being thrust deep into the mesoglea. Although Cnidaria constitute by far the most readily available food for Bullia, other carrion, of higher calorific value, is preferred when present. The tunicate Pyura and bivalves such as Choromytilus, Aulacomya and, on Muizenberg Beach, Schizodesma, are commonly washed up during storms and these are avidly eaten (Brown, 1961a, 1964d, 1981). Dead, slightly decayed Pyura appears to be preferred to fresh material and constitutes the best bait for attracting large numbers of the whelks. Even a small piece of Pyura, attached to a piece of string and dragged along the wet sand will commonly result in hundreds of whelks emerging from the sand and surf and congregating on the faint line left by the passage of the bait (Brown, 1961a). Bullia has not been seen to graze plant material and pays no attention to plant matter in the laboratory. This does not, however, preclude the possibility that very young whelks may take plant as well as animal food. Bullia will, on occasions, turn active predator and attack other members of the sandybeach community. Gilchrist (1916) observed either B. digitalis or B. rhodostoma consuming living amphipods in such numbers that he thought this to be their normal diet, although in reality such predation usually follows a long period of starvation (Brown, 1964d). The whelks will also prey on living polychaete worms, the bivalve Donax, and even on the prawn Callianassa, where these are available. When attacking small prey, the whelk crawls over it and then folds its foot around the animal; the proboscis is then introduced into the fold from the side and the prey eaten. On Koeberg Beach (a continuation of that at Ou Skip) there are areas where the sandy shore is invaded by rocky outcrops which support a psammophilic rocky-shore fauna, notably the mussel, Choromytilus meridionalis. Bullia digitalis has been observed in large numbers surfing up the sandy slope onto these rocks, where any gaping Choromytilus are eagerly consumed (Brown, unpubl.). Cannabalism has, however, not been witnessed under any circumstances, including prolonged starvation in the laboratory, nor are other species of Bullia eaten, dead or alive. McGwynne (1980) found the supply of carrion to Eastern Cape beaches to be highly erratic. It was composed mainly of coelenterate medusae and siphonophores, the summer input of medusae ranging from 0.15 to nearly 4 kg·km−1·h−1 wet weight. The winter input

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was so small that no stranded carrion at all was seen between July and October. Under these circumstances it is not surprising that Bullia is an opportunistic feeder and that it can consume food up to the equivalent of a third of its own tissue weight in a single meal (Brown, 1961a). Such a meal may last 10 min when the food consists of jelly-fish but longer when the food is tougher. The calorific value of the meal may be sufficient for 14 days of active existence without the whelk feeding again, when the food has a high calorific content (Brown, 1981).

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Fig. 5.—Dorsal view of dissected radular mass of Bullia rhodostoma, with oesophagus removed (from P.Du Preez, unpubl.): 1, pharynx; 2, proboscis cavity; 3, radula; 4, bolsters; 5, radula sheath; 6, portion of circular muscle around radula mass; 7, anterior radula protractor muscle; 8, left bolster attached to circular muscles; 9, posterior region of bolster attached to lateral radula retractor muscle; 10, posterior radula protractor muscle; 11, radula retractor muscle; 12, ventral protractor muscle of radula mass; 13, radula sac; 14, right lateral radula retractor muscle; 15, central radula retractor muscle; 16, radula sac retractor muscle; 17, roots of left lateral radula retractor muscle; 18, proboscis retractor muscles; 19, septum; 20, anterior protractor muscle of radula mass; 21, muscle bulb.

McGwynne’s (1980) calculations for the calorific values of jelly-fish, however, suggest that a meal of medusa may last the whelk little more than two days. The speed and efficiency with which Bullia ingests its food might imply the presence of special adaptations. In principle, however, the feeding apparatus is the same as that described for Nassarius obsoletus by Brown (1969). Bullia radulae have been studied by Peile (1937), Barnard (1959), and Meredith & Brown (in press). They are typical rachiglossan radulae with three sharp, cusped teeth per row and without any special features which might distinguish them from those of other nassarid whelks. As in the Nassariidae as a whole, the elaboration of the proboscis allows deep penetration into animal tissues and there is an oesophageal valve which permits protrusion and elongation of the proboscis without regurgitation of food (P.Du Preez, unpubl.). The fully extended proboscis might at first appear to be of extreme length (up to times the length of the shell) but in fact some other whelks have equally long proboscides (Pearse & Thorson, 1967). Indeed, the only feature which may distinguish Bullia and be related to its feeding efficiency is the robustness of the musculature associated with the feeding apparatus (see Fig. 5). The proboscis wall itself is highly muscular and has longitudinal, circular and diagonal muscle layers. It is of the pleurembolic type characteristic of the Buccinacea. In this type of proboscis only the basal portion is able to invaginate and when it does so it draws the distal part into a cylindrical fold of the body wall, the proboscis sheath. In many pleurembolic proboscides, the basal region is folded into a permanent introvert, even when the proboscis is extended. In Bullia the proboscis is, however, fully everted on extension (P.Du Preez, unpubl.). The alimentary canal is relatively short, as one would expect in a carnivorous whelk. Assimilation efficiency in Bullia is unknown but is probably high. Measured values for assimilation efficiency in carnivorous marine gastropods vary from 52% to 79% (Huebner & Edwards, 1981). Brown (1971b) has analysed the sequence of events which leads to feeding in B. rhodostoma and B. digitalis. The first response is due to distance perception by means of olfaction, substances such as trimethylamine, emanating from the food, stimulating the osphradium (Brown & Noble, 1960). When not recently fed, the whelks react by emerging from the sand if buried and, when not already in the wash zone, surfing ashore.

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Once in the wash, the animal crawls in the direction of the food, the relative straightness of the path taken depending on the degree of interference from the wash. The siphon is held horizontally, so that water from the surface film overlying the sand continues to be sampled and passed over the osphradium when the wash has retreated. Thrusting movements of the proboscis are initiated on contact with the potential food and the whelk may also feel the food with its cephalic tentacles. The thrusting movements of the proboscis soon cease if the object is not potential food. In order for feeding to take place, the object must contain substances which add a further stimulus, possibly acting on the sub-radula organ. The volatile amines and other substances which served to attract the whelk do not suffice to initiate feeding. Amino acids are, however, effective in this respect, while sugars are not. Brown (1961a) demonstrated that the shallow-water, non-surfing B. laevissima would only feed under water and then only in relatively non-turbulent conditions. In contrast, B. digitalis (at Hout Bay, on the west coast of the Cape Peninsula) would feed under water if the opportunity presented itself but, because turbulence usually precluded this, food was normally consumed in or immediately above the wash zone. This has subsequently also been found to be largely true of other west coast beaches. B. digitalis, however, does not have to compete with other species of Bullia on these beaches. On the south and east coasts of South Africa, where such competition commonly occurs, the behaviour of B. digitalis may be different. McGwynne (1980) states that, on the beaches she studied in the Eastern Cape, it was B. rhodostoma which dominated access to carrion stranded in the wash zone, B. digitalis (and B.pura) taking food on its path through the surf. Judging from the experience of Brown (1961a), this would only be possible where wave action and turbulence are not too severe. Observations on Muizenberg Beach, False Bay (Brown, unpubl.) show that both B. digitalis and B. rhodostoma may feed in the wash zone, the species tending to be separated by a long-shore segregation rather than a vertical zonation; nevertheless, when carrion is attacked on its way through the surf, it is B. digitalis which is usually involved, B. rhodostoma waiting until the food is carried to the shore. As McGwynne (1980) points out, the heavier shell of B. digitalis lends it greater stability and thus more opportunity to feed under turbulent conditions. It has also been noted on Muizenberg Beach that the highest carrion, at the very top of the wash zone tends to be eaten by B. rhodostoma rather than B. digitalis. The tendency of the Bullia species to form aggregations on some beaches has been attributed largely to sorting of the animals by waves and currents (Brown, 1971b) but there can be no doubt that their feeding habits tend to maintain these groups. B. melanoides on Shertallai Beach, in southwestern India, also forms aggregations and this too appears to be associated with feeding (Ansell et al., 1972). PREDATORS OF BULLIA Sandy beach communities are exposed alternately to two sets of predators as the tide rises and falls. Those predators invading the intertidal sands from the land when the tide is low include small mammals, reptiles, and birds, the former two groups mainly at night, while the birds are mainly active during the day. Predatory fish and crabs invade the area as the tide rises, the fishes being most active around dawn and dusk.

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McLachlan et al. (1979b, 1981) reported Sanderlings (Crocethia alba) feeding on Bullia rhodostoma on Eastern Cape beaches, although this has not been observed on Cape Peninsula beaches. Indeed, after many years of observation on the latter beaches, it can be stated that predation of Bullia by birds is negligible (Brown, 1964d, 1971b). The rôle of small mammals and reptiles is less easy to assess due to the relative infrequency of night observation. There are, however, a number of important aquatic predators of South African Bullia, including the holocephalan, Callorhynchus capensis (Brown, 1971b), elasmobranch fishes such as Rhinobatis, the Batoidea and others (Brown, 1964d, 1971b; McLachlan et al., 1981) and the teleosts Coracinus capensis, Lithognathus lithognathus, and Rhabdosargus holubi (Brown, 1964d; McLachlan et al., 1981). Caine (1974) found that the crab Ovalipes will sometimes eat molluscs and McLachlan et al. (1979a) noted that Ovalipes punctatus preys on Bullia rhodostoma on Eastern Cape beaches. Du Preez (1981) has made a detailed study of predation by Ovalipes punctatus on Bullia and the bivalve Donax on these beaches. Crabs from King’s Beach had stomach contents with fragments of shells from Donax sordidus (52%) and Bullia pura (20%), while those from Maitland River Beach contained Donax sordidus (38%), D. serra (29%) and Bullia pura (16%). B. rhodostoma, which is more abundant than B. pura, is less available to the crabs because of its higher position on the shore. Ovalipes uses a variety of methods to obtain the flesh of the whelks. Small Bullia are usually completely crushed by the chelae, while larger individuals are crushed to various degrees. The flesh of large whelks may be extracted through the mouth of the shell without the shell being crushed and sometimes whole animals are extracted in this way. In a series of choice experiments, Ovalipes from King’s Beach showed a preference for Donax serra over Bullia rhodostoma but accepted B. pura in preference to both B. rhodostoma and Donax. In contrast, crabs from Maitland River Beach preferred Bullia rhodostoma to B. pura or Donax. Du Preez (1981) comes to the conclusion that Ovalipes selects the species to which it has least access in the field. Small to medium-sized Bullia were preferred to large individuals, although the mean size of the B. rhodostoma chosen increased with increasing size of crab. Ovalipes punctatus is probably the most important predator of Bullia on Eastern Cape beaches and possibly also on other South African beaches. An interesting observation is that neither B. rhodostoma nor B. digitalis appear to be eaten by the voracious isopod Eurydice longicornis, which is to be found on the same Cape Peninsula beaches, although it will tear flesh from still-living Bullia laevissima. B. laevissima, however, is not attacked by Exosphaeroma truncatitelson, an isopod commonly occurring in the same ecosystem (Brown, 1973). From time to time, empty Bullia shells may be found on the beach, each with a hole bored through it, usually in the second or third whorl. This could be due to a boring naticid whelk, although as the hole is straight-sided and not recessed, some other molluscan predator could be responsible. Palmer (1980) has noted that no less than 95% of the empty shells of Nassarius kraussianus in the Bushman’s River estuary had holes bored by Natica tecta and suggests that one Natica may consume, on average, one adult Nassarius per week. The frequency of bored Bullia shells, however, has always been low except in Hout Bay during the late 1950s when the Bay was subject to severe fish-factory

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pollution (Brown, unpubl.). The beaches became strewn with empty shells, a high percentage of which had been bored. Empty Bullia shells are commonly used for protection by hermit crabs such as Diogenes brevirostris, and Barnard (1963) found such a crab sharing a shell with a stillliving Bullia. Unlike many psammophilic forms (Donax, Arenicola, Callianassa, etc), sandy-beach whelks have not been much exploited by man, either for bait or food. During 1981 it was found, however, that vast numbers of Bullia were being removed from Muizenberg, with the object of using their shells for ornamental purposes (M.Picker, pers. comm.). POPULATION STUDIES, BIOMASS, CHEMICAL COMPOSITION Bally (1981) has investigated the ecology of three sandy beaches on South Africa’s west coast and confirms the highly uneven long-shore distribution of the macrofauna, including that of Bullia digitalis. He found that this species made up between 0.5 and 16% of the macrofaunal biomass at Melkbosstrand (near Ou Skip) and up to 23% at Yserfontein, further north. Isolated individuals of B. laevissima were also found at Yserfontein on two occasions. Maximum biomass values for B.digitalis were 29.6 g·m−1 of beach at Melkbostrand and 48.3g·m−1 at Yserfontein, both on the basis of decalcified dry tissue weight. Bullia was found to be rare at Rocherpan, the third beach investigated, making no significant contribution to the biomass. Quantitative sampling of both macrofauna and meiofauna has been carried out intensively on four Eastern Cape beaches in the vicinity of Algoa Bay. These four beaches display an extreme range of macrofaunal biomass, largely resulting from greatly differing food supplies (McLachlan, 1977b). B. rhodostoma proved to be an important member of the fauna on all four beaches, while on Maitland River Beach this species was accompanied by B. digitalis and lesser numbers of B.diluta and B. pura. On beaches in Sardinia Bay, B. rhodostoma accounted for over 90% of the total macrofaunal biomass, while at St George’s Strand it made up only 5.5%. On Maitland River Beach it accounted for <1%. Vast populations of the bivalve Donax serra occur, however, on the latter two beaches and form the bulk of the biomass (McLachlan, 1977b). In Algoa Bay the young of Bullia rhodostoma hatch as crawling juveniles between December and February, reaching a length of 10 mm after about a year and some 40 mm after 10 years. Animals with a shell length of >50 mm are probably about 20 years old. The von Bertalanffy growth equation is (McLachlan et al., 1979a). The annual mortality on King’s Beach was calculated as 0.79 but on Maitland River Beach was only 0.34 during the period of study (McLachlan & Van der Horst, 1979). The main cause of “natural mortality”, as opposed to predation, is the effect of storms which carry animals above the driftline, where they die of exposure (Brown, 1971b; McLachlan et al., 1981), for they cannot crawl on or burrow into sand which is not saturated with water (Brown, 1961a). The loss of production through all causes of mortality is about 10% of the available production (McLachlan et al., 1981). The mean decalcified dry biomass (B) of B. rhodostoma on King’s Beach was 209mgm−2 and production by growth (P) was calculated as 189 mg·m−2·yr−1, giving a P/B ratio

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of 0.9. This figure agrees well with that for Maitland River Beach. Most production by adults (of a shell length >15mm) goes into reproduction, particularly in female individuals, and production by reproduction is estimated as 135mg·m−2·yr−1 on King’s Beach (McLachlan et al., 1979a). On Maitland River Beach the P/B ratio is 0.06. A wide fluctuation in calorific values was found, without any obvious seasonal trend except for higher values just before spawning. On King’s Beach the average calorific value was 4.54 kcal·g−1 (19.04 kJ·g−1 ±1.55), rising to about 4.95 kcal (20.75 kJ) in the November pre-spawning period. On Maitland River Beach, the average calorific value was 4.50 kcal·g−1 (18.84 kJ·g−1) dry tissue weight (McLachlan & Van der Horst, 1979). These figures correspond well with those for Cape Peninsula B. rhodostoma and with those of the B. digitalis population at Ou Skip, for which Brown (unpubl.) has calculated an average value of 4.52 kcal·g−1 dry tissue wt. Brown (1971b) attempted to assess the size of the adult population of B. rhodostoma on Muizenberg Beach, False Bay, by a capture-recapture method. He estimated that there were about 600 individuals per 100 m stretch of beach or, in terms of biomass, some 5.8 g·m−1 on a wet wt. basis (about 0.97 g decalcified dry tissue wt). B. digitalis occurs in roughly the same numbers on this beach, while B. pura is much less common. Total Bullia biomass on the beach is about 12 g·m−1 wet wt (2 g dry wt), accounting for some 28% of the total macrofaunal biomass. There is no doubt that it would make up a lesser percentage were not Donax and Callianassa heavily exploited on this beach, an exploitation which has not generally extended to Bullia. On Shertallai Beach, in southwestern India, where B. melanoides is the only representative of the genus, the highest densities were recorded in January and October; up to 330 or 340 animals per metre stretch of beach, or 53 g·m−1 wet tissue wt (15.35 g·m−1 dry tissue wt) (Ansell et al., 1972). Mean biomass was 11.62 g·m−1 wet wt (3.37 g·m−1 dry wt), representing some 17.44 kcal·m−1. As in the South African Bullia species, calorific values of B. melanoides proved to be quite variable, ranging from 3.9 to 4.5 kcal·g−1 dry tissue wt. These values are lower than those obtained from bivalve molluscs inhabiting the same beach (Ansell, Sivadas & Narayanan, 1973). The P/B value for Shertallai B. melanoides is given as 13.5 (Ansell, McLusky, Stirling & Trevallion, 1978) but this value may be somewhat too high (see Brown, 1981). Size distribution data for this species suggest two periods of recruitment a year (Ansell et al., 1972), a situation similar to that reported for B. vittata on India’s eastern coast (Crichton, 1942b), but differing from the South African species, where only one period of recruitment a year is indicated. Crichton suggests that in B. vittata “full growth is attained within a period of not more than six months”. It is not clear what this statement means for a species showing continuous growth but it is highly likely that both this species and B. melanoides grow faster and have shorter life-spans than the South African species (Brown et al., 1978). In all Bullia populations investigated, females outnumber males and attain a much larger size (Brown, 1971b, McLachlan et al., 1979a, McGwynne, 1980). A preponderance of females is, in fact, found in most prosobranch populations (Fretter & Graham, 1962) and it is also usual for females to reach a larger size than the males (Webber, 1977), although not always to such a marked degree as in Bullia. Studies of the chemical composition of B. melanoides from Shertallai Beach and of B.

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digitalis from Ou Skip have been undertaken by Ansell et al. (1973) and Miss F.M.Da Silva (unpubl.), respectively. The latter worker found that adult B. digitalis, sampled in spring and early summer, had some 70% protein, calculated as a percentage of the dry tissue wt, and only 5.5% lipid. These values were consistent and remained so in the laboratory. Figures for total carbohydrates and for free reducing sugars were more variable, the former averaging 9% and the latter 3.7% of the dry tissue wt. These values declined steadily in the laboratory in unfed animals. These figures are not inconsistent with those obtained by Ansell et al. (1973) for B. melanoides over a period of a year. Crude protein averaged 66% of the dry tissue wt (S.D.±3.94), lipids 5% (S.D.±1.0) and carbohydrates 2% (S.D.±0.7). There is an indication of a seasonal trend in the results, the highest values for lipid and total nitrogen being reached in October (immediately after the monsoon season) and the lowest values in April and May (before the start of the monsoon). It is not clear whether this reflects seasonal variation in the availability of food or seasonal changes in the gonads. Seasonal changes in weight of a standard animal show the same trend. McGwynne (1980) has found similar weight changes in B. rhodostoma from the southern coast of South Africa and has suggested that they are correlated with the breeding cycle.

ENVIRONMENTAL TOLERANCE LIMITS TEMPERATURE Brown (1961a) considered briefly the upper temperature tolerances of Bullia digitalis and B. laevissima collected from Cape Peninsula sandy beaches. He placed his animals in beakers of sea water held in water baths, which were then heated at a rate of about 1°C in 5 min to the required test temperature. The water was imediately allowed to cool to 19°C and the whelks left for 2 h to see how many of them recovered. By repeating the procedure a number of times at various temperatures, he assessed the upper temperature causing 50% mortality as about 39°C for B. digitalis and 35.5°C for B. laevissima. He related this difference to the difference in distribution between the two species, and particularly as between the intertidal and subtidal modes of life. McGwynne (1980, also reported in Ansell & McLachlan, 1980) has performed similar, although not identical, experiments on all three intertidal species on Eastern Cape beaches. She found the thermal death point to be 38°C for B.rhodostoma, 35°C for B. pura, and only 33°C for B. digitalis. These results certainly reflect the differences in zonation of the species in the area from which they were collected. It has already been mentioned that in the Eastern Cape B. digitalis does not normally leave the water, while on Cape Peninsula beaches it is frequently exposed to air (Brown, 1971b). Whether this difference accounts entirely for the different results gained by Brown (39°C) and McGwynne (33°C) is doubtful and it seems likely that differing methodology and perhaps differences in the criteria used to assess death, have made a significant contribution to the disparity. Ansell & McLachlan (1980) have carried out more sophisticated temperature tolerance studies on B. rhodostoma in the Eastern Cape. They used temperatures between 26 and

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38 °C, with a difference of 2°C between each group of animals, and observed the whelks after 1, 3, 6, 12, 24, 48, 72, and 96 h at the test temperature. Large individuals showed a greater thermal tolerance (31.2°C) than small individuals (29.9°C) but the figures in general agree well with the results of similar studies on Nassarius reticulatus and Nassa pygmaea from the Mediterranean and the north Atlantic (see Table I). It is also of interest that Bullia rhodostoma has a higher thermal tolerance than the more deeply buried bivalves of the genus Donax from the same beaches (Ansell & McLachlan, 1980).

TABLE I Comparison of LT50 values for southern warm-temperate, European warm-temperate and North Atlantic species of Nassariidae (after Ansell & McLachlan, 1980)

Area South Africa (southern coast)

Mediterranean

North Atlantic 1 2

Species

Mean 48–72 h LT50

Bullia rhodostoma1 (large)

31.2

B. rhodostoma1 (small)

29.9

Cyclonassa neritea2

33.0

C. donovani2

32.0

Nassa pygmaea1

31.0

Nassarius reticulatus2

30.5

N. reticulatus2

28.0

Acclimated to 20°C. Acclimated to 15°C.

SALINITY Krijgsman & Brown (1960) and Brown (1961a) touched on the upper and lower salinity tolerances of Bullia digitalis and B. laevissima from Cape Peninsula beaches, while McGwynne (1980) studied the salinity tolerances of B. digitalis, B. rhodostoma, and B. pura from the Eastern Cape. The methods used were not identical. Brown (1961a) added either distilled water or concentrated sea water at set intervals of time and attempted to discover the concentrations at which crawling effectively ceased. He found that B. laevissima had difficulty in crawling below a salinity of 23‰, while the intertidal B. digitalis could crawl in salinities down to 18‰, despite obvious swelling and abnormal turgour of the foot. Both species crawled in high salinities up to 45‰. McGwynne (1980) found that normal crawling in her three species was possible between salinities of 14 and 56‰ and that the animals survived for more than 12 h in these salinities. B. rhodostoma survived a salinity of 60‰ for 12 h, while B. digitalis and B. pura did not. While the relatively unsophisticated methodology of both Brown and McGwynne render the results hardly worthy of detailed study, it may be concluded that the subtidal B. laevissima is the least tolerant of salinity changes, while B. rhodostoma,

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which migrates highest up the shore, appears to be the most tolerant, although only slightly more so than B. digitalis and B. pura. An interesting observation made by McGwynne (1980) is that the jets of water expelled by retracting whelks after subjection to high or low salinities for more than 4 h conformed in salinity to that of the surrounding water. Brown (1964b) has shown that the water making up the jets is drawn chiefly from the free space and the mantle cavity, and has suggested that in the expanded animal there may be a free interchange of water between these spaces. McGwynne’s observations provide additional evidence in favour of this view. On the other hand, Brown & Meredith (1981) have indicated that in B. digitalis from the Ou Skip population, the respiratory current passing through the siphon into the mantle cavity ceases or is markedly reduced in high or low salinities. These authors suggest that this may help to protect the pallial organs, and particularly the gill, from osmotic stress. Such protection does not, however, seem to apply to the Eastern Cape whelks studied by McGwynne, except perhaps for periods of <4 h. WATER LOSS Brown (1961a) studied the rate of water loss in air, the degree of loss which proves fatal, and the tolerance of various saturation deficiencies in B. laevissima and B. digitalis. The whelks were caused to retract into their shells and dried before the experiments were done. It was found that the rate of water loss of the two species, expressed as a percentage of wet tissue wt, and measured at 19°C in calcium chloride desiccators, was not very different, although B. digitalis initially lost water rather faster than did B. laevissima. B. digitalis was shown, however, to be far more tolerant of water loss than B. laevissima. Thus the period for which desiccation can be withstood is much longer in B. digitalis. It was also shown that B. digitalis survived various degrees of saturation deficiency longer than B. laevissima, this being most marked at the higher humidities. McGwynne (1980) performed similar experiments on the intertidal Eastern Cape species. As with the other tolerance studies reported above, the fact that the methods used were not identical to those employed by Brown (1961a), makes direct comparison of the results impossible. McGwynne showed that retracted B. digitalis in general loses water faster than B. pura, and this species faster than B. rhodostoma, under conditions of constant humidity (and presumably temperature) (Fig. 6). She also demonstrated that small individuals of B. rhodostoma become desiccated much faster than large individuals. B. rhodostoma survived all the test humidities for significantly longer than did the other two species. Thus, as in the assessment of temperature and salinity tolerances, B. rhodostoma appears to be better adapted to fairly long periods out of water than either B. pura or B. digitalis, this being associated with its generally higher position on the shore. B. laevissima seems to be less tolerant of wide fluctuations in conditions than any of the three intertidal species investigated. It should, however, be noted that the desiccation experiments undertaken both by Brown and by McGwynne involved extremely artificial conditions in the laboratory; in particular, the tests were carried out on retracted animals, whereas in the field retraction into the shell is rare, the animals remaining expanded for days or weeks. When in air the expanded whelk carries a considerable amount of water in its free space, mantle cavity, and aquiferous system (Brown, 1964b), water which is lost

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on retraction. The actual data gained in the above experiments are thus virtually meaningless except for comparative purposes and should even then be treated with caution.

REPRODUCTION As is typical of prosobranch gastropods, the sexes are separate in nassarid whelks and fertilization is internal, the male possessing an elongate penis. A rudimentary penis may also be present in the female, for example in Bullia

Fig. 6.—Rates of water loss as a percentage of initial wet tissue weight in three Eastern Cape Bullia in dry air: vertical bars represent standard deviations; average shell length of the B. digitalis used was 28 mm, B. pura 22 mm, and B. rhodostoma 34 mm; (from McGwynne, 1980).

rhodostoma (McGwynne, 1980), but the male penis is easily distinguished by the extensive coiling of the ejaculatory duct and its thick muscular wall. REPRODUCTIVE CYCLES

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Breeding in the Nassariidae normally takes place in late spring or summer, although some members of the genus Nassarius may breed in mid-winter (Pechenik, 1978). Members of the Bullia group are essentially summer breeders (Crichton, 1942a; Penchaszadeh, 1971b; McLachlan et al., 1979a; McGwynne, 1980). Breeding is, however, not necessarily synchronous throughout the range of the species and it appears that conditions in some areas lead to copulation taking place in autumn (Brown, 1971b; J.Omer-Cooper, pers. comm.) although the eggs in such cases may only hatch the following summer. In both males and females of Bullia, the gonads are inextricably intermingled with the digestive gland, making the determination of a gonosomatic index difficult or impossible. Histological examination of the gonads by McGwynne (1980) and G.Van der Horst (pers. comm.) has shown, however, that the reproductive cycles of B. rhodostoma, B. digitalis, and B. pura on Eastern Cape beaches are similar, although not identical. Gametogenesis in B. pura begins in April and lasts for about three months until July, vitellogenesis occurring in August and September. The eggs are spawned in November and December. In B. rhodostoma and B. digitalis, gametogenesis occurs between March and May, vitellogenesis and egg storage taking place from June or July until October or November in B. rhodostoma and until December or January in B. digitalis. Spawning then begins.

COPULATION Copulation in B. digitalis and B. rhodostoma was described by Brown (1971b), while copulation in the Indian B. vittata was witnessed on a single occasion by Crichton (1942a). Crichton noted that the expanded foot of one individual was wrapped closely and entirely around the shell of a second, larger animal. The foot of the second individual was also expanded and was engaged in maintaining the position of the animal in the sand, against a strong current. Similar copulatory positions occur in B. rhodostoma and B. digitalis, the partners facing the same direction, the foot of the male tightly grasping the shell of the female, being drawn back at one point to allow entry of the penis into the female’s shell opening. The proboscis of one of the partners, and sometimes both, may be extruded and used to gnaw at the foot or head of the other whelk. In B. rhodostoma, such copulating pairs have been found mainly below the surface of the sand but in B. digitalis they have most commonly been observed in the surf, the partners frequently being washed backwards and forwards by the waves. Copulation may last for at least half an hour (Brown, 1971b). Both B. rhodostoma and B. digitalis display a second method of copulation, a few centimetres below the sand surface; the animals lie side by side, facing opposite directions, with the right side of the male’s foot overlapping that of the female so that the shells almost touch, allowing the penis to be inserted (Brown, 1971b). Copulating pairs of B. pura were noted by McGwynne (1980) during August and September on Eastern Cape beaches. She found copulation in B. rhodostoma and B. digitalis to occur during September-October and November-December, respectively. Unfortunately McGwynne does not describe the copulatory processes in any of the species. The spermatozoa are apparently stored by the female between copulation and spawning. McGwynne (1980) has found that on Eastern Cape beaches the young of both B.

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digitalis and B. pura occur only offshore and that breeding females of both species migrate from the intertidal zone into deeper, less turbulent water to deposit their egg cases. This occurred during January and February in B. digitalis and between February and April in B. pura. Some of the Indian species of Bullia may also migrate offshore to breed or when the young are ready to hatch (Crichton, 1942b; Ansell et al., 1972). THE EGG CAPSULES The Nassariidae appear to have roughly equal numbers of species with pelagic and nonpelagic larval development. Moreover, there seems to be an evolutionary relationship between egg-capsule morphology and type of development, as has previously been shown for the Columbellidae (Bandel, 1974). First, there are species which produce tough vase- or flask-shaped egg cases, usually containing >30 eggs which hatch as pelagic, planktotrophic veligers. The shapes vary from the spiny multi-faceted capsules of Nassarius obsoletus (Scheltema, 1962) to the four-sided vase of N. tiarula (Houston, 1978) and to the flattened vase of N. reticulatus (Lebour, 1931). A second major group of nassarid species produce triangular cases, each containing a single egg which undergoes demersal, direct or ovoviviparous development. Intermediate between these two groups are the species of Tritia studied by Amio (1956, 1959). The Bullia group exhibits trends which are somewhat different from other Nassariidae. In those species of Buccinanops studied by Penchaszadeh (1971a), the young hatch as shelled, crawling juveniles from egg capsules attached to the callous region of the maternal shell. Nurse eggs outnumber developing eggs to a considerable degree. A similar situation is found in Dorsanum moniliferum (Penchaszadeh, 1971b), the egg cases being attached to the maternal shell by short peduncles. The greatest number of such capsules recorded on a single shell was 21. In each capsule only a single egg develops but it also contains up to 1600 nurse eggs. Nurse eggs appear to be characteristic of a number of neogastropod families and most egg cases are filled with a nutritive albuminous fluid (Anderson, 1960). Within the genus Bullia, egg cases are known from B. tennis, B. rhodostoma, B. digitalis, B. livida, B. tranquebarica, and B. melanoides. Four egg capsules of B.tennis were dredged from False Bay during the University of Cape Town’s Ecological Survey in 1952 and were subsequently described by Barnard (1959). They were “thin and membranous, soft, transparent, with an attachment thread at each end, one of the threads being less coiled than the other on all four capsules”. Barnard’s figure shows an irregular oval purse containing a single, shelled individual apparently nearly ready to emerge as a crawling juvenile. The dimensions of the case are not given but the shell of the embryo measured 5.3×3.5 mm. Three of the four egg cases contained only a single individual, the fourth containing “numerous eggs”. The egg cases of B. digitalis and B. rhodostoma resemble those described for B. tennis in being flattened, oblong purses but have generally been found to be more regular in outline than that figured by Barnard (1959) possibly because they were fresher. The case typically measures about 2×1.2 cm and contains >1500 eggs arranged in clumps of 50 to 100, although one capsule of B. rhodostoma containing only 24 eggs has been recorded (Brown, 1971b). The single capsules are deposited from 4 to 12 cm below the surface of

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the sand, below the level of low water of springs. Although the egg cases are not attached to the parent or to any other object, there is some evidence that the maternal whelk may stay with its single brood for some period (Brown, 1971b). Parental care is not unknown among gastropods and Hipponyx, as well as some Calyptraeidae, are known to brood over eggs laid external to the body (Yonge, 1953; Ansell & Trevallion, 1970). The tropical Indian Bullia melanoides produces a number of capsules at one time, each of which resembles those of the South African Bullia species (Ansell & Trevallion, 1970). The capsules are at first oval in outline, with two slender, coiling filaments, one of which is more coiled than the other. Later the outline becomes irregular as the walls begin to collapse, due most probably to the utilization of the fluid contents by the developing young. The capsules are held beneath the ventral surface of the maternal foot, with the filaments of adjacent capsules twisted together. Each capsule contains between 1 and 5 embryos, together with a number of spherical bodies (?nurse eggs). In female B. melanoides carrying eggs, the edges of the foot are curled over centrally to form a tube which protects the capsules (Fig. 7). Crawling can still take place, however, only the edges of the foot being used, while the capsules are held clear of the substratum. In B. livida and B. tranquebarica from the Madras area of India, the egg capsules are also apparently held on the ventral surface of the foot. Crichton (1942a) described these egg cases as being “in the form of oval transparent capsules, each about six millimetres long, held together by mucus matter”. The Senegal species B. miran is, in contrast, said to produce vase-shaped egg cases, the eggs emerging as planktotrophic veligers (G.Thorson, in Ansell & Trevallion, 1970). It seems probable that this species has been wrongly placed systematically and does not in fact belong to the genus Bullia. LARVAL SUPPRESSION Radwin & Chamberlin (1973) and Shuto (1974) have suggested that certain neogastropods, in becoming specialized to particular habitats, have found it advantageous to suppress the planktonic larval stages so that the young, emerging as crawling juveniles, settle predominantly in the parental environment. This results in a higher percentage survival than would otherwise be achieved, a survival which would be further enhanced in Bullia by the maternal care of the egg-cases, which are either attached to the female or brooded over by her. Presumably the advantages attendant on these methods outweigh the loss of wide dispersal by means of veligers. In this regard it is of interest to note that the ovoviviparous Nassarius kraussianus releases the young as very late veligers which soon settle (Palmer, 1980), while N. plicatus may release either late veligers or crawling juveniles (R.N.Kilburn, pers. comm.). In any case the suppression of planktonic larvae may not be quite as limiting for dispersal as might at first be thought. For example, Bullia rhodostoma, which is essentially a south coast South African species, extending west to Cape Point, is to be found again in the mouth of Saldanha Bay, after a gap of some 150 km (Brown, 1977). The cost of reproduction to Bullia remains unknown, but it may be noted that reproductive effort in the Mollusca is higher in semelparous species than in those which are iteroparous. In the latter, reproductive effort increases with successive breeding seasons (Browne & Russell-Hunter, 1978).

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The appearance of a penis-like structure in females of B. rhodostoma suggests the possibility of protandric hermaphroditism, although there appears to be no other evidence to support this view (McGwynne, 1980). It is, perhaps, more likely that the rudimentary penis of the female indicates an hermaphroditic ancestry and indeed several authors have suggested that protandry might be the primitive condition for the molluscan phylum as a whole.

BLOOD CIRCULATION, FLUID SPACES, AND EXTENSION OF THE FOOT THE BLOOD SYSTEM Brown (1964b) studied the anatomy of the blood system and used the radio-opaque dye Thorotrast to follow the movements of the blood (haemolymph)

Fig. 7.—Two individuals of Bullia melanoides from Shertallai Beach in

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southwestern India: both are carrying egg cases attached to the sole of the foot, which is curled over to form a brood pouch; the individual with the foot slightly open, revealing the egg capsules, carried 33 cases, while the other retained only three; (from Ansell & Trevallion, 1970).

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in Bullia digitalis and B. laevissima. He found the general anatomy of the system to be not dissimilar to that of Buccinum (see Fig. 8). Both dorsal and ventral buccal sinuses are present, draining into a ventral cephalic sinus which links up with the extensive pedal sinus via a dorsal cephalic sinus and an oesophageal system of subdivided sinuses. The transverse septum, dividing the anterior sinuses from the posterior ones is welldeveloped. There are two routes by which blood may leave the pedal sinus system; a large cephalopedal vein drains the central part of the pedal sinus, while two pallial collecting vessels drain the left side and carry blood round the mantle collar to a pallial sinus. From here blood can enter the pallial system of vessels or pass to the gill. The dorsal pallial vessel enters the visceral sinus system, providing direct vascular communication between the anterior and posterior regions of the body.

Fig. 8.—Diagram of Bullia with foot extended, showing the principle features of the blood system (from Trueman & Brown, 1976; after Brown, 1964b).

Blood passing up the cephalopedal vein meets blood coming from the visceral sinus system via a visceral vein, both streams draining into the kidney vessels. An alternative

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route is via a pallial anastomosis which links the afferent renal vein to the branchial and pallial systems. Blood returning to the heart must thus pass through either the kidney or the gill. The vascular system is, however, so constituted that not all the blood leaving a sinus system must of necessity pass through the heart before re-circulating. The cephalopedal and visceral veins in fact form a continuous tube of fairly uniform bore, without valves, linking the visceral and pedal sinus systems. It can be demonstrated that blood passes up this tube from the pedal to the visceral sinuses when the animal retracts into its shell and it is assumed that the reverse route is taken during extension of the foot, the heart, gill, and kidney all being short-circuited. Any dangerous build-up of pressure appears to be avoided by the presence of the pallial anastomosis already mentioned. EXTENSION OF THE FOOT Brown (1964b) assumed that the passage of blood from the visceral to the pedal sinus accounted entirely for the extension of the foot and that the turgour of the expanded foot was brought about by the tonus of the visceral somatic musculature acting against the contraction of the pedal muscles. He admitted, however, that the visceral musculature appeared to be too poorly developed to support this hypothesis. Trueman & Brown (1976) found that, although a flow of blood into the pedal sinus is necessary for the inflation of the foot, actual extension is achieved by the extension of the columellar muscle. This muscle originates from the columella in the second whorl of the shell and its fibres radiate to all parts of the foot but particularly to the propodium and the dorsal surface of the metapodium, where many fibres are inserted onto the operculum. The fibres of this muscle form a three-dimensional network similar to that seen in squid mantle. These fibres appear to antagonize each other directly, without the participation of any fluid or hard skeleton. Thus not only will contraction of the longitudinal fibres draw the foot into the shell but contraction of the transverse and/or dorso-ventral fibres, accompanied by relaxation of the longitudinal fibres, must result in extension of the foot. Sections of columellar and opercular muscles from a number of other prosobranch gastropods reveal the same three-dimensional network of fibres and indicate that this mechanism of foot extension is wide-spread (Brown, in prep.). It is possible that different neuro-transmitters are involved in stimulating the longitudinal fibres and those at rightangles to them (Trueman, pers. comm.), a possibility which is to be investigated experimentally. A network of muscles has been found to occur in the neck region of Bullia, such that their contraction might effectively prevent the flow of blood from pedal to visceral sinus systems (Trueman & Brown, 1976). These neck fibres are considered more likely to provide the necessary antagonism for turgour of the foot than is the relatively weak visceral somatic musculature. Even if the pedal and visceral systems are effectively isolated from one another while the foot is expanded, the question remains whether the heart pumps blood through the pedal sinus under these circumstances. Brown (1964b) claimed that such circulation was in evidence, but this requires confirmation. The question may be an important one in view of the fact that Bullia commonly maintains its foot in the expanded condition for days or even weeks. Moreover, if circulation through the foot is interrupted or restricted during activity, then the pedal musculature may be

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dependent on oxygen diffusing directly through the wall of the foot, a possibility which is at present being investigated (Brown, in prep.). BLOOD VOLUMES One of the most obvious features of whelks of the genus Bullia is the exceptionally wide expansion of the foot. Although the foot is very thin, this does not compensate for its breadth, so that the foot volume is higher than that in the rocky-shore prosobranchs with which it has been compared (Brown, 1964b). This, together with the fact that the interior of the foot contains relatively little tissue (Trueman & Brown, 1976), might lead one to predict an unusually high blood volume for the animal. Brown (1964b), using a dyedilution method, assessed the blood as making up about 65% of the total wet tissue volume of B. digitalis and B. rhodostoma, while blood accounted for some 57% of B. laevissima. These figures, while high, may not be exceptional for prosobranchs. SEA-WATER SPACES During the first half of the nineteenth century, many workers believed that the expansion of various structures in marine Mollusca, especially the foot, was brought about by the uptake of sea water. Later work on a number of species revealed that, in fact, these structures were dilated by means of blood withdrawn from other parts of the body. The view thereafter tended to be that water never played any part in expansion of the foot and accounted for no part of the body volume (Hyman, 1967). Bernard (1968), RussellHunter & Apley (1968), and Russell-Hunter & Russell-Hunter (1968) were, however, able to demonstrate clearly and quantitatively that sea water plays a major rôle in the expansion of the foot of the naticid snail Polinices. When Bullia retracts into its shell, a jet of fluid issues from the hind margin of the foot, between the posterior cirri. On rapid contraction, two lateral jets are also apparent. Lewis (1911), assuming correctly that these jets consisted of sea water, studied histological sections of the foot and claimed to have found lateral and posterior pores at the points of origin of the jets, leading from the pedal sinus to the exterior of the animal. Gilchrist (1916) also believed that the foot of Bullia was inflated by means of sea water. This assumption was also made by Brown (1961a) but was later corrected by Brown & Turner (1962), who found that the main space within the foot, the pedal sinus, did not communicate with the exterior at any point. There were, indeed, small aquiferous spaces in the posterior region of the foot which opened directly to the ventral surface, but these were not large enough to account for any considerable fraction of the water expelled as jets. These aquiferous spaces (Fig. 9), like the much more extensive spaces in Polinices, have possibly originated as an elaboration of pedal mucous glands. Brown & Turner (1962) were able to identify a pair of grooves on the ventral surface of the foot, running from the openings of the aquiferous spaces to the mid-point of the posterior margin of the foot. In life, the common opening of these grooves gives the appearance of a pore from which the posterior jet originates. By introducing a length of arterial catheter through the siphon, it is possible to draw off most of the water in the mantle cavity and to measure its volume. When the whelk is then

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caused to retract into its shell, the associated jets of water are reduced in volume by approximately one third (Brown & Turner, 1962). Brown (1964b) was able to confirm that sea water from the mantle cavity contributed to the jets, by introducing Thorotrast into the cavity and taking radiographs of the jetting whelk. He was also able to demonstrate a contribution from the free space which develops between the shell and the visceral hump when the foot is extended. There is evidence of exchange of water between the mantle cavity and the free space. Water from the mantle cavity and free space is conducted to the edge of the foot by means of flaps of skin and it is from the ventral edges of these that the anterolateral jets originate during rapid retraction. During less violent, normal retraction, the water is conducted from the anterolateral flaps backwards through temporary channels formed by a curling over of the extreme lateral margins of the foot, so that this water contributes to the posterior jet (Brown & Turner, 1962). Despite the relatively large amount of water contained in the free space, longitudinal sections of the shell, as well as radiographs, show that there is no resorption of the inner whorls to increase the size of the cavity, as occurs in some gastropods (Brown, unpubl.). While retraction takes place both under water and in air, in response to appropriate stimuli, emergence from the shell only occurs under water. Brown (1964b) suggested that Bullia will not under any circumstances draw air into the free space; it seems more likely that it is the aquiferous spaces that must fill with water rather than air. In Bullia, as in the large naticids (Russell-Hunter & Russell-Hunter, 1968), the animals do not readily retract into their shells and remain continuously expanded for many days or even weeks. This is not associated with the surfing habit, as non-surfing species of Bullia display the same phenomenon. A retracted Bullia can, however, achieve full expansion of the foot in seconds, whereas in the naticid Polinices duplicatus expansion takes from 3–8 min, and even longer in Lunatia heros (Russell-Hunter & Russell-Hunter, 1968). These differences are probably linked with the extent of the aquiferous systems.

LOCOMOTION SURFING Three forms of locomotion are apparent in sandy-beach whelks—surfing, crawling, and burrowing. Surfing used to be referred to as “passive transport” in the surf (Brown, 1961a, Ansell & Trevallion, 1969). In reality it is a highly active process with a relatively high energy cost per unit time (Brown, 1979a), although cheap in terms of distance travelled (Brown, 1982). It is, however, unlikely that the whelk has any control over the direction in which it is carried. Surfing involves extreme turgour of the foot, combined with twisting and waving movements which are often vigorous. These movements begin in response to water currents (Brown, 1961a, 1971b; Ansell & Trevallion, 1969). Trueman & Brown (1976) introduced a thin plastic cannula through the dorsal surface of the foot and into the pedal haemocoel of Bullia digitalis and connected it to a pressure transducer linked to a George Washington pen-writing oscillograph. Haemocoelic pressure was monitored while the animals lay on the surface of the sand in an

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experimental tank and the water gently stirred by blowing across its surface. Even such mild currents led immediately to a marked increase in pressure associated with an increase in pedal turgidity. Indeed the resting pressure of some 2×10−2N·cm−2 can be increased instantaneously to a value of 20×10−2 N·cm−2 in response to such currents. This cannot be due to additional haemolymph passing into the

Fig. 9.—Longitudinal section through the posterior region of the foot of Bullia digitalis, showing aquiferous space: AS, aquiferous space; PH, pedal haemocoel; (photograph by Brown & Turner, unpubl.).

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Fig. 10.—An individual of Bullia rhodostoma commencing to burrow: the characteristic track left by the animal while crawling over wet sand is clearly seen; (from Brown, 1971b).

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foot nor to an increase in circulatory pressure, in the absence of any signs of heart pulse pressures in the pedal sinus. It is most likely that the markedly increased pressure is due to tension being developed in the pedal muscles. CRAWLING Crawling in Bullia has been studied by Brown (1971b) and particularly by Trueman & Brown (1976), who used ciné film and electronic recording to analyse the process. Cilia are not used in locomotion, as they are in many prosobranchs, including Polinices (Trueman, 1968). Crawling is achieved by alternate movement and anchorage of the propodium and metapodium. With the posterior region of the foot anchored, the propodium is raised slightly and thrust forwards. At the same time the cephalic shield advances and is raised. In the second phase, the propodium becomes anchored anteriorly, allowing the shell and metapodium to be drawn forewards while the cephalic shield moves down and backwards. Anchorage of the propodium is achieved largely by lateral swelling so that, viewed from above, crawling resembles to some extent the breast-stroke used by swimmers. The track left by the crawling whelk is characteristic, the outline of the propodium being apparent in the sand at regular intervals (about every 0.5 cm for a large Bullia digitalis) (Fig. 10). Extension of the propodium must involve relaxation of the longitudinal muscles anteriorly, tension in the transverse fibres to maintain constant pedal width, and the application of the force of contraction of other muscles through the hydraulic system. This results in a stretching of the pedal sole anteriorly, in the same way as occurs at the commencement of any retrograde wave, as for example in the anterior segments of an earthworm (Trueman, 1975). Instead of travelling progressively down the foot, however, as happens in the locomotory wave of Patella (Jones & Trueman, 1970), extension of the propodium is followed by its anchorage and the contraction of longitudinal pedal muscles and anterior fibres of the columellar muscle to pull the metapodium and the shell, respectively, forwards in a step-like movement. A single pressure pulse is apparent in the pedal haemocoel during each locomotory cycle, coincident with pedal extension (Trueman & Brown, 1976). Such pulses do not exceed 4×10−2N·cm−2 and are usually much lower (i.e., less than a fifth of the value associated with the response to currents). Although crawling in Bullia does not depend on adhesion to the substratum, mucusproducing cells are present in the sole of the foot and this mucus can act as an adhesive when necessary, allowing the animal to crawl up the side of a glass tank or plastic bucket (Brown, 1971b). Adhesion is of importance during copulation and may also play a rôle during feeding, particularly under water in turbulent conditions, when the whelk clings to the food rather than the substratum.

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BURROWING Burrowing has been studied more intensively than either surfing or crawling (Brown, 1961a, 1971b; Ansell & Trevallion, 1969; Trueman & Brown, 1975, 1976). Burrowing is essentially an adaptation of normal surface locomotion, the propodium being inserted into the sand as a mobile, freely progressing wedge (Brown, 1971b). Ansell & Trevallion (1969) measured pressure changes produced in the sand by freely burrowing B. melanoides. The tension exerted by the columellar muscle as it pulls the shell forward and downward through the sand was also measured, using an isometric myograph attached to the apex of the shell. They were able to confirm that in burrowing the same movements are apparent as in crawling. They contrasted the mechanism of burrowing in B. melanoides with that in Oliva gibbosa, a whelk found in the same area. Trueman & Brown (1976) carried this work further, using Bullia digitalis as their main experimental animal but also making observations on B. rhodostoma and B. laevissima. In addition to measuring pressure pulses in the sand near a burrowing animal, they monitored pressure changes in the pedal haemocoel, in the mantle cavity, and in the free space. They also recorded the forces exerted by the whelk, by tethering it from the apex of its shell to a force transducer. The forces used in burrowing are much greater than those recorded during surface locomotion; by a factor of×20 in respect of tensile force exerted on the shell and×7 in respect of pressure pulses in the pedal haemocoel. The greater force exerted by the fibres of the columellar muscle inserted into the propodium is a measure of the increased anchorage of the foot when buried, which allows the shell to be pulled forward more powerfully. A composite diagram of the events associated with a single digging cycle in B. digitalis is presented in Figure 11. There are two pressure pulses per cycle in the pedal haemocoel, the first effecting propodial extension, the second, larger pulse occurring as the shell is drawn forwards. The latter phase is also associated with single pressure peaks in the free space and the mantle cavity. Penetration of the propodium is most probably aided by the simultaneous lifting of the cephalic shield, resulting in water being drawn into the space beneath the shield. It was also observed that water was ejected from the mantle cavity during burrowing. This phenomenon was confirmed by introducing methylene blue solution through the siphon of a burrowing whelk; blue water was then ejected during the next digging cycle. The ejection of mantle cavity water corresponds to movement of the shell forwards, or the development of tension in the columellar muscle, and recordings of pressure increases in the sand adjacent to the animal were apparent at the same time. The function of this water ejection is almost certainly to liquefy the sand so as to facilitate forward movement of the shell. The mantle cavity pressure is at a maximum when water is expelled, while the shape of these pulses is in general similar to those recorded directly in the sand. The mantle has little musculature, so that the expulsion of water is probably due to contraction of the columellar muscle compressing the mantle cavity. The pressure in the cavity tends to fall below atmospheric pressure after the water has been ejected, presumably as water is drawn in by its re-expansion. The pulses in the free space occur after ejection of water from the mantle cavity but are probably caused by the same agency. As far as burrowing is concerned, Bullia exhibits a remarkable degree of

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convergence with burrowing bivalve molluscs. Not only is the pattern of alternating anchorages similar but both eject water from the mantle cavity to liquefy the sand in front of the shell.

Fig. 11.—Analysis of the activity of Bullia digitalis during a single digging

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cycle, reconstructed from ciné film and electronic recordings, showing the relationship between pressure (×10−2 N·cm−2), tensile force and movement: force was measured by a thread from the apex of the shell to a force transducer, maximal values being obtained when the shell was drawn forwards, minimal when the propodium was being extended, as indicated by the movement curves (---, shell;——, metapodium;—·—, propodium); water ejection from the mantle cavity and the raising of the cephalic shield are also indicated; (from Trueman & Brown, 1976).

Brown (1961a) has shown that the particle size distribution of the sand influences the rate of burrowing and that B. digitalis is unable to burrow into sediments consisting of particles >3.2 mm in diameter. In the laboratory. medium-sized B. digitalis from Ou Skip, burrowing into sand from the same site, completed the process in an average of 10 digging cycles lasting a total of 45s, the animal ceasing burrowing when the shell was covered by sand with only the tip of the siphon protruding above it (Trueman & Brown, 1976). Bullia is capable of burrowing not only under water but also into saturated sand exposed to air. It fails, however, to burrow into dry sand or even into wet sand if the water table lies below the surface (Brown, 1961a). In the wash zone of a sandy beach, Bullia tends to burrow while covered with water; the animals often commence burrowing in the wash and then lie immobile as the water retreats and the sand becomes more compact, completing the burrowing process during the next wash (Brown, unpubl.). EMERGENCE FROM THE SAND Emergence from the sand, in response to water currents, food or other factors, usually takes place more or less vertically, the lateral margins of the foot pushing downwards and forcing the shell up through the sand. Normal crawling movements then free the animal completely when necessary (Brown, 1961a, 1971b; Ansell & Trevallion, 1969).

RESPIRATION, METABOLISM, AND ENERGY UTILIZATION OXYGEN CONSUMPTION Estimates of metabolic rate, as evidenced by rates of oxygen consumption, have been measured for Bullia melanoides on tropical Indian beaches and for B. digitalis, B. rhodostoma, and B. pura on South African beaches. B. melanoides at local sea temperature (30°C) proved to have a rate of oxygen uptake about an order of magnitude higher than that of the temperate B. digitalis at 15°C, measured at similar and controlled levels of activity (Brown, Ansell & Trevallion, 1978). This is consistent with the findings of Ansell et al. (1978) that productivity on tropical sandy beaches is about ten times higher than that of temperate beaches. It also suggests an absence of temperature acclimation, a suspicion confirmed for B. digitalis at Ou Skip by Brown & Da Silva (1979) and for Eastern Cape populations by Dye & McGwynne (1980). Brown & Da Silva were also able to demonstrate a markedly flattened rate-temperature curve for the

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species (Fig. 12), such that an increase in temperature within the limits likely to be encountered by the whelk in the field does not lead to a significant increase in the rate of oxygen consumption (Q10=1.1). This relative temperature independence was measured at a fixed routine rate of activity and not simply during quiescence, which has been established for a number of marine organisms (Newell, 1979). This greatly depressed response to temperature increase may well be of importance to an animal which must conserve energy in temperatures that may range from 8 to 17°C in a short time (Brown & Jarman, 1978; Branch, Newell & Brown, 1979). Differences in metabolic rate between B. melanoides, B. digitalis and whelks such as Nassarius reticulatus. which has a still lower rate than B.

Fig. 12.—Acute rate-temperature curve for large adults from the Ou Skip population of Bullia digitalis suspended in a constant current: all values have been transformed to a standard-sized animal of 750 mg dry tissue weight; vertical bars represent 1 S.D. on either side of the mean values; (after Brown & Da Silva, 1979).

digitalis at ambient temperatures, may be associated with different life-styles, involving not only different levels of activity but also different life-spans and reproductive cycles (Brown et al., 1978). Much of the consistency gained in the above work was due to the successful maintenance of a controlled activity level during measurement of oxygen uptake. To achieve this, each individual was suspended in sea water from the apex of its shell, the water being stirred by means of a magnetic stirrer. The currents thus produced not only mixed the water and created a continuous flow over the electrode used to monitor oxygen depletion, but also stimulated the animal to expand its foot and wave it from side to side, a response to currents also observed in the field (Brown, 1961a). Brown (1979a) was able to demonstrate that the speed at which the water was stirred did not significantly affect the rate of oxygen uptake of the whelk providing it was held constant. On the other hand, when the speed of the stirrer was changed continuously, so as to provide fluctuating surges of turbulence, much higher levels of oxygen uptake were measured, the animals

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responding actively to changes in current velocity. The values obtained under these conditions may approximate those associated with transport in the surf. Dye & McGwynne (1980), in studying the oxygen consumption of B. rhodostoma, B. digitalis, and B. pura from Eastern Cape beaches, did not attempt to standardize the activity levels of their animals. The whelks were placed in containers of sea water without sand and without being restrained in any way. McGwynne (1980) reports that under these experimental conditions “the animals had the foot fully extended and slow movements were observed for most of the time”. Dye & McGwynne confirmed that medium-sized B. digitalis have a flat rate-temperature curve within the environmental temperature range, while the curve for small individuals was somewhat less flat. They confirmed also the low exponent value for oxygen uptake against tissue weight obtained by Brown & Da Silva (1979) (0.589) but showed that B. rhodostoma had an exponent close to the generalized value of 0.75. B. pura appeared to have less regulation of its metabolic rate over the temperature range than did other species, having higher and more variable Q10 values. B. pura showed considerable counter-clockwise rotation of the ratetemperature curve in winter, a feature shown to a much smaller extent by B. rhodostoma and not at all by B. digitalis. ENERGY COSTS OF ACTIVITIES Trueman & Brown (1976), using a force transducer to measure the mechanical work performed by Bullia during crawling and burrowing, assessed the cost to the animal by assuming a gross efficiency of 20%, as previous workers had done for a variety of animals. Brown (1979b), however, measured the actual uptake of oxygen during burrowing, simultaneously obtaining readings of the forces exerted by the animal (see Fig. 13).

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Fig. 13.—Method of measuring the energy cost and efficiency of burrowing in Bullia, as described by Brown (1979b): the animal is placed in a tank of sea water, protected from the air by a layer of medicinal paraffin, over a sandy substratum; it is tethered by the apex of its shell to a force transducer at an angle of some 45°; this induces burrowing movements, the force of which is recorded on a pen recorder; at the same time water samples are removed for oxygen analysis; uptake of untethered, immobile animals may be compared with that of burrowing animals to determine the cost of burrowing activity.

He obtained much higher figures for the gross energy cost of burrowing (2.1×10−2 cal over and above maintenance costs for a burrowing event of 10 digging cycles lasting 45 s) and assessed the efficiency of the process as only 6%. In the absence of appropriate data from other prosobranch gastropods, the cost of crawling and burrowing in Bullia may be compared with that of crawling in the terrestrial slugs studied by Denny (1980). The metabolic cost of crawling in these slugs is given as 215 cal·kg−1·m−1, considerably more than that reported for other forms of locomotion in other animals. In contrast, the cost of burrowing in Bullia is only approximately 60 cal·kg−1·m−1, and the cost of crawling 36 cal, a cost which is comparable with that of running in terrestrial animals (Schmidt-Nielson, 1972). This difference between a slug’s crawling and crawling in Bullia is most probably to be attributed largely to the fact that the slug progresses by adhesive crawling and expends much energy producing the mucus by which it adheres to the substratum, while Bullia crawls in a series of “breast-strokes”, without the necessity for adhesion (Trueman & Brown, 1975, 1976). The energetic costs of the various activities of B. digitalis from Ou Skip, other than feeding, have been listed by Brown (1979c), who assumed an oxycalorific value of 3.4 cal per mg O2 consumed by the whelk. Newell and Brown found some difficulty in quantifying the changes in oxygen uptake associated with feeding (see Brown, 1979d). Since then, Brown & Meredith (unpubl.) have repeated the feeding experiments conducted by Crisp, Davenport & Shumway (1978) on Nassarius reticulatus. Their results for Bullia differ from those gained with Nassarius in that the oxygen consumption associated with feeding proves to be no greater than that associated with crawling, even in starved individuals. Stimulation by substances emanating from the food causes a negligible increase in oxygen uptake. AN ACTIVITY (TIME-ENERGY) BUDGET Brown (1981, 1982) has observed Bullia digitalis in the field for many years and attempted to follow single individuals from their time of emergence from the sand on a falling tide until their final burrowing as the tide rose. In each case the length of time spent surfing, crawling, and feeding was noted, as well as the number of burrowing cycles. Where crawling was observed, the distance moved was also often recorded. The most obvious feature of these data is the considerable variation in the amount of time spent by different individuals in different activities during the tidal cycle, and particularly in the amount of time spent above the sand (Brown, 1981). Nevertheless, a time-energy

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(activity) budget has been prepared for adult, non-feeding B. digitalis on the beach at Ou Skip (Brown, 1982) and this is reproduced in Table II.

TABLE II Average energy expenditure by eight non-feeding female Bullia digitalis on Ou Skip beach during single tidal cycles of activity, referred to a standard-sized animal of 750 mg dry tissue weight (from Brown, 1981)

Activity

Oxygen uptake per min (µg)

Actual time Oxygen uptake per (min) period (µg)

Energetic equivalent (cal)

Transport in surf

20.8

14.5

302

1.03

Crawling

11.3

23.0

260

0.88

Burrowing

18.8

7.5

141

0.48

Emerging

18.8

3.8

72

0.25

Buried (observed)

9.3

42.0

391

1.33

Buried (residual)

10.6

653.2

6924

23.54

744

8090

27.51

Total for tidal cycle

Two categories of buried state are listed in this table: buried (observed) and buried (residual). The former refers to observed and timed periods of burial between crawling and/or surfing. Buried (residual) refers to the assumed state of the animal for the unobserved portion of the tidal cycle, while the tide was in. This is by far the longest of the activity states and it is here that the greatest errors in calculation are likely to be. Brown (1979a) reported a mean oxygen uptake of 560 µg·h−1 for animals of 750 mg dry tissue wt buried in sand under moving water (=initial rate in stagnant water). It is thought that this figure is too low to reflect field conditions, as it can be shown in the laboratory that buried whelks take a step down into the sand in response to simulated wave crash (sea water poured into their tank from a beaker). Thus to the oxygen uptake of buried animals in the laboratory has been added the oxygen consumption associated with two burrowing steps per min, giving a total of 634 µg O2·h−1. That animals in the most active state consume oxygen at approximately twice the rate of resting animals conforms well with observations on other molluscs (Newell, 1970) and on marine invertebrates in general (Newell, 1979). That the most active state should be surfing, previously referred to as “passive transport in the surf”, is, however, somewhat surprising. During surfing the animal displays maximum turgour of the foot and waves it vigorously from side to side, often with violent twisting movements; presumably it is the combination of these actions which demands high energy expenditure. Nevertheless, in

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terms of distance travelled—up to 10 m or more in a few seconds—surf transport is far less costly than crawling. Indeed it compares favourably with gentle walking in terrestrial vertebrates (Schmidt-Nielson, 1972). Burrowing is comparable with surfing as far as energy demand per unit time is concerned, but is by far the most costly activity in terms of distance travelled (Brown, 1982). COST OF FREE EXISTENCE The average cost of free existence for large B. digitalis reaching the shore (wash zone) and crawling but not feeding, works out at 27.51 cal for a 744-min tidal cycle. As it is considered unusual, at Ou Skip at least, for large adults to be active in the surf for two consecutive tidal cycles and as only about 12% of the animals come ashore on any one cycle, the average energy cost for an animal of 750 mg dry tissue wt is calculated as 52 cal in 24 h (Brown, 1981). The extremes of energy expenditure for such a standard animal in 24 h are 51.9 cal, for a continuously buried animal, and 61.6 cal, for a whelk which is extremely active during two consecutive tidal cycles. In view of the animal’s essentially flat rate-temperature curve (Brown & Da Silva, 1979), it is quite unnecessary to adjust these figures for fluctuating field temperatures or for the season of the year; nor does feeding activity significantly alter the calculation. One of the conclusions to emerge from this study is the low cost to the animal of being active during the tidal cycle, as opposed to remaining buried. In homeothermic animals the cost of free existence has been calculated to be about double that of the resting or standard rate (Meon, 1973) but in B. digitalis the cost of free existence is only about 1.14 times the best estimate of standard rate (Brown, 1981). As all the data discussed above were transformed to those of a standardsized animal of 750 mg dry tissue wt, we might assess the cost of free existence of any given individual from the Ou Skip population by applying the formula

where W is the dry tissue wt of the whelk in question in mg and 0.598 is the regression coefficient of weight against oxygen consumption determined by Brown & Da Silva (1979). This resolves into

for a 24-h period. Two problems arise, however, in the application of this formula. First, the regression coefficient of 0.598 as determined by Brown & Da Silva (1979), using their own data as well as those of Brown et al. (1978), is lower than the generally accepted exponent value of 0.75 for organisms as a whole (Zeuthen, 1953; Hemmingsen, 1950, 1960). One might argue that it is, in fact, more appropriate to use this higher figure rather than that obtained

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experimentally. This is what Ansell et al. (1978) have done in calculating the population respiration of the Indian B. melanoides, although the exponent actually obtained experimentally for this species was only about 0.5 (Brown et al., 1978). As Dye & McGwynne (1980) have confirmed the low coefficient for a different population of B. digitalis, there seems little merit in adopting a different, if more universal, figure. It may also be noted that Houlihan, Innes & Dey (1981) have used an exponent value of 0.65 for several gastropod species, while Huebner & Edwards (1981) have calculated and used an exponent of 0.536 for Polinices duplicatus. A second problem in the extrapolation of the data to all size classes in the Ou Skip population is that young individuals are more active than the large animals used in all the respiratory and energetic experiments. Newly hatched whelks appear to be active in the surf during virtually every tide, while large animals may be active on only 12% of tides (Brown, 1971b). When this is taken into account, the regression of weight against energy expenditure will be lowered still further. It is not clear, however, whether a curve of degree of activity against weight would show a logarithmic relationship or not. If it did then the final exponent could be as low as 0.5. For the time being, therefore, it seems reasonable to adopt the figure of 0.598 (perhaps approximated to 0.6). Not only have Ansell et al. (1978) used a theoretical exponent of 0.75 but they have extrapolated to their population of B. melanoides as a whole measurements of oxygen consumption obtained while the animals were continuously active, thereby obtaining figures which are too high to reflect actual population respiration in the field. Nevertheless, this does not invalidate their conclusion that population respiration and productivity on tropical beaches is about an order of magnitude higher than on temperate beaches. The activity budget constructed for the Ou Skip population of B. digitalis (Brown, 1981, 1982) cannot be transferred confidently to other beaches with-out modification. Brown (1981) has stressed the considerable variation in the activity period between different beaches and has linked these differences to differences in beach slope and other factors. McLachlan et al. (1979e) and McGwynne (1980) have demonstrated that, during their periods of observation, the centre of gravity of the B. digitalis population on Maitland River Beach moved up and down the shore continuously with the tides. Virtually continuous activity on the part of individual animals is implied. This might result in an average cost of free existence for a standard animal of some 63 cal in 24 h, instead of the 52 cal calculated for the Ou Skip population. EFFECT OF SALINITY ON OXYGEN UPTAKE The experiments which led to the drawing up of Table II and the estimates of free existence for B. digitalis, were all conducted at a temperature of 15°C and in sea water of a salinity approximating 34‰. B. digitalis may, however, sometimes be found on intertidal sands at the mouths of estuaries (Brown, 1961a) and it may also encounter reduced salinities on non-estuarine beaches during and after rain (Brown, 1971a). Brown & Meredith (1981) thus investigated the oxygen uptake of suspended animals at several different salinities. It was found that both elevated and reduced salinities evoked a marked decrease in oxygen uptake (Fig. 14). This is in contrast to what one

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Fig. 14.—Mean oxygen consumption per hour at 15°C, of large, unfed, female Bullia digitalis in a range of salinities, transformed to that of a standard-sized animal of 750 mg dry tissue weight (from Brown & Meredith, 1981).

might expect from reviews of the literature (Vernberg & Vernberg, 1972; Newell, 1979), although Shumway (1979), working with 11 species of gastropods, reported withdrawal into the shell as a primary response to decreased salinity, oxygen uptake ceasing thereafter. Oxygen uptake in Bullia also virtually ceases after retraction (Brown, unpubl.) but the animal shows no tendency to retract into its shell in response to any but the most severe and sudden salinity changes. Brown & Meredith (1981) have shown, however, that the normal respiratory current of Bullia either stops or is markedly reduced at both high and low salinity and attribute the reduced oxygen consumption largely to this fact, together with a decrease in activity. The cessation of this current may be seen as protecting the pallial organs, including the gill, from osmotic stress, at least for a limited period. ANAEROBIOSIS The respiratory current also ceases or is reduced when Bullia is buried in sand under stagnant sea water for a prolonged period and this, too, is associated with a decrease in oxygen consumption, very low rates being reached after some time (Brown, 1979a). It must be questioned whether, under these circumstances, there is not a shift towards anaerobic metabolism. Other evidence for anaerobiosis includes the fact that the whelks can survive for at least 18 h in oxygen-free sea water without apparent harm (Brown, 1964c), while Miss F.M.Da Silva (in Brown, 1979b) has kept individuals of B. digitalis in sealed jars of sea water for five days and found the animals still alive, although

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severely stressed, after that period. The identification of an end-product of anaerobiosis proved difficult initially (Brown, 1979b) but recent work indicates that a major endproduct is probably alanine (Brown, unpubl.). This finding is consistent with results from a number of other molluscs (Newell, 1979). It is not clear whether anaerobic metabolism in Bullia pertains only to certain conditions, such as prolonged quiescence or anoxia, or whether there also continues to be an anaerobic component at other times. Such activities as burrowing do not appear to result in an oxygen debt, but this could simply mean that the end-products of anaerobic pathways are not being re-utilized, at least not immediately. The answer to the problem of assessing the extent of anaerobiosis during the various activities of the animal would undoubtedly be the use of sophisticated direct calorimetry to measure total heat output. While direct calorimetry is now being applied to marine invertebrates (Pamatmat, 1979; J.M.Shick, pers. comm.), the types of apparatus currently available do not permit of measurements being taken during specific activities such as burrowing, crawling or surfing. Brown (in press) has measured the rate of oxygen diffusion through the pedal wall of B. digitalis and has shown that when the expanded whelk is surrounded by oxygensaturated sea water, more oxygen diffuses into the foot than is required by the entire animal, even during its most vigorous activities. Under these circumstances the gill may be seen as functioning chiefly to supply oxygen to the visceral mass, particularly while the blood supply to and from the pedal haemocoel is restricted by contraction of the neck muscles, as it appears to be during locomotion and possibly during other activities (Trueman & Brown, 1976). The possibility arises of the visceral mass being forced towards anaerobic respiration when the respiratory current is reduced, while the pedal musculature enjoys an excess of oxygen. When the whelk is buried in sand, however, the water in contact with the foot may soon become depleted of oxygen and the animal may then have to rely almost solely on the gill and mantle to supply oxygen brought from above the sand surface in the respiratory current flowing through the siphon. Work at present in progress at the University of Cape Town includes the measurement of oxygen and carbon dioxide concentrations in the mantle water and in the pedal haemolymph of B. digitalis under a variety of controlled conditions, as well as the characteristics of the haemocyanin. The respiration of whole-animal homogenates is also being studied, using a Gilson respirometer. Initial results from the latter experiments indicate that homogenates of B. digitalis have a Q10 >2, in contrast to the Q10 of only 1.1 determined for intact animals. This finding, however, requires confirmation. ATP CONCENTRATIONS Brown & Gedye (in press) have demonstrated wide variations in adenosine 5′triphosphate (ATP) levels and adenylate energy charge (AEC) in natural populations of B. digitalis. AEC values only a little over 0.5 have been recorded frequently, particularly in whelks which have been very active in the surf. Such animals returned unfed to the laboratory and allowed to remain buried in sand, however, consistently restore their ATP levels, AEC values approaching the maximum value of 1 after about 24 h. This restoration of ATP from adenosine 5′-diphosphate (ADP) is brought about at the expense

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of carbohydrate reserves (F.M.Da Silva, unpubl.), probably via arginine phosphate. The fact that the AEC is allowed to decline in active animals to values which must be considered critical for the well-being of the whelk, raises the possibility that glycolysis or the associated storage processes involving arginine phosphate may not be able to supply energy fast enough for such activity, so that ATP is broken down faster than it can be produced. This would result in an “ATP debt”, which will also represent a small oxygen debt if the ATP is restored aerobically, although not of the type associated with vertebrate muscle. Theoretically the ATP debt incurred during a particular activity should be added to the cost of that activity as measured by oxygen uptake or direct calorimetry. A simple calculation, however, demonstrates that the cost of restoring this ATP is negligible. The average ATP concentration in B. digitalis is 5.72 mg·g−1 dry tissue wt or 4.29 mg ATP for a standard animal of 750 mg dry wt (Brown & Gedye, in press). If half of this were converted to ADP during activity, and given that one mole of ATP costs the animal 17.5 kcal, the debt would represent only the merest fraction of a calorie. On the other hand, although the tissue concentration of ATP is relatively low (0.5% of the dry tissue wt), the rate of turnover is high. It can be calculated that a standard animal utilizing 52 cal in 24 h produces 1.5 g of ATP from ADP during this period, an amount equal to about 40% of its wet tissue wt. This may seem extraordinarily high until we remember that the human body produces virtually its own weight of ATP (≈70 kg) in 24 h (Karlson, 1968). ATP concentrations in Bullia fall well within the range of the bivalves studied by Ansell (1977). Active bivalves, capable of sudden bursts of energy, have higher concentrations of ATP than have relatively inactive species and this is almost certainly true of prosobranch gastropods as well. It may be noted that in the foot of Bullia ATP values may approach 1% of the dry tissue wt, with an average of about 0.85%, while the visceral mass has much lower concentrations (Brown & Gedye, in press).

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Fig. 15.—Scanning electron micrograph of part of the surface of an osphradial leaflet of Bullia digitalis, showing tufted ciliated cells and secretory cells: membrane-bound secretory vesicles are apparent, adhering to the surface of the epithelium; scale bar, 10 µm; (photograph by P.F.Newell).

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SENSORY SYSTEMS Bullia, unlike Dorsanum and most other Nassariidae, lacks eyes and it has not been possible to demonstrate any response to light or shadow either in the field or in the laboratory (Brown, 1971b). The animal also appears to be insensitive to sound waves. Under these circumstances, the chemical senses and touch play major rôles in the life of the whelks. Gibbons (1878) noted the sensitivity to touch of Bullia rhodostoma and assumed an “acute sense of smell” to be present to compensate for the lack of eyes. Brown & Noble (1960), using surgical and behavioural techniques, demonstrated that the osphradium was the olfactory organ, a finding since confirmed for other marine gastropods (Bailey & Laverack, 1963, 1966). CHEMORECEPTION The osphradium of B. digitalis is an elongate, bipinnate organ situated on the floor of the mantle cavity near the gill and immediately below the opening of the siphon. The osphradial ganglion forms its axis and from this axis arise sensory leaflets on either side. The base of each leaflet is also attached along one edge to the floor of the mantle cavity. In live preparations, motile cilia can be seen on the lateral margins and on the surface of the leaflets. The fine structure of this osphradium has been described in some detail by Newell & Brown (1977), using both transmission and scanning electron microscopy. The organ does not differ markedly from the prosobranch osphradia described by Dakin (1912), Welsch & Storch (1969) and Crisp (1973), although innovated, tufted ciliated cells (Fig. 15) are described for the first time. It is possible that these cells have both afferent and efferent synapses (Newell & Brown, 1977). The cilia of these cells have well developed striated rootlets and the cells are packed with mitochondria. They may well be the cells concerned with chemoreception. Several other types of ciliated cells also occur. There are typical ciliated cells, bearing long cilia and occurring in dense zones, as well as cells bearing less well developed cilia and identical in appearance to the S3 cells of Welsch & Storch (1969) described from Buccinum. The cuboid supporting cells appear to be secretory, and secretory vesicles may be observed in section, pushing through the apical mat of convoluted microvilli. Mucussecreting cells are scattered throughout the surface of the leaflet, while the Type 4 cells of Welsch & Storch and the “onion cells” of Crisp (1973) occur below the surface epithelium. In freshly-dissected Bullia the osphradium is very darkly pigmented, the pigment being most probably melanin, and in section melanophores may be seen scattered throughout the surface. The osphradium of Bullia has been shown to be extremely sensitive to trimethylamine,

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the animals responding positively to concentrations of under 0.01 mg/1 (Brown, 1961b, 1971b). The whelks also respond to substances such as skatole and indole by emerging from the sand, although their orientation to the source of these chemicals is poor. High concentrations of indole, while detected, stimulate avoidance behaviour in the form of deeper burrowing. Solutions of γ-buterobetaine also evoke avoidance, although a combination of this substance with trimethylamine proves attractive in certain proportions (Brown, 1971b). Substances such as trimethylamine, while serving to attract the animals and inducing them to touch the potential food and to make thrusting movements with the proboscis, do not initiate feeding, and artificial food is rejected unless it contains amino acids. These may well be detected by the subradula organ (Brown, 1971b), although this organ appears to be diffuse and not well developed in Bullia (A.Gedye, pers. comm.). The cephalic tentacles may also be sensitive to chemical stimuli, as the whelk commonly touches the food with them before feeding. The posterior cirri, at the extreme hind end of the foot, appear to have a sensory function, the animal testing the water with them before emerging from its shell. It will not emerge when the water is polluted with substances such as hydrogen sulphide (Brown, 1964b,d). Early suggestions that these cirri may act as rudders during surfing are incorrect. P.F.Newell & Brown (unpubl.) have found what appear to be well-developed tactile receptors on the surface of the cirri. The posterior cirri are absent in Dorsanum. The entire upper surface of the foot of Bullia possesses a general chemical sensitivity (Brown, 1961a) and this appears to be most acute around its margins. A. Hodgson (in prep.) has identified sensory cells on the pedal surface which are unlike any described previously from the Mollusca. MECHANORECEPTION Brown (1971b) reported the presence of a pair of statocysts in B. digitalis and B. rhodostoma, lying on either side of the pedal ganglion and resembling those described for Buccinum by Dakin (1912). The shell itself, however, acts as a giant statolith, forces acting on it probably being sensed through receptors in the columellar muscle. It seems likely that the detection offerees acting on the shell may at least partly account for the whelk’s ability to assess the strength and direction of water currents while it is on the surface of the sand. Nevertheless, the animal can detect currents when buried with only the tip of its siphon protruding above the sand and will respond to such currents by emerging actively.

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Fig. 16.—Lateral views of the circumenteric nerve ring of Bullia digitalis (after D.Donnelly, unpubl.). A, left side: 1, cerebral commissure; 2, fused cerebral and pleural ganglia; 3, siphonal nerve; 4, cerebro-pedal connective; 5, nerve to ventral body wall; 6, pedal ganglion; 7, 8, nerves to posterior and middle regions of foot; 9, 10, nerves to anterior region of foot; 11, to lateral body wall and proboscis retractors; 12, to head; 13, tentacular nerve; 14, buccal ganglion; 15–17, left oesophageal and pharyngeal nerves; 18, 19, oral nerves. B, right side: 1, cerebral commissure; 2, tentacular nerve; 3–5, oral nerves; 6, 7, oesophageal nerves; 8, buccal ganglion; 9, cerebropedal connective; 10, pedal ganglion; 11, anterior pedal nerve; 12, to head and mantle; 13, anterior pedal nerves; 14, nerves to posterior and middle regions of foot: 15, to lateral body wall; 16, to floor of body cavity; 17, to lateral body wall and proboscis retractors; 18, right visceral connective; 19, to ventral body wall; 20, columellar nerve; 21, subintestinal ganglion; 22, supra-intestinal ganglion; 23, to anterior regions of osphradium and gill; 24, left visceral loop connective; 25, to posterior osphradium and gill; 26, fused cerebral and plueral ganglia.

THE NERVOUS SYSTEM The nervous systems of B. rhodostoma and B. digitalis have formed the subjects of student projects undertaken by G.Read and D.Donnelly, respectively (unpubl.). There appear to be only trivial differences in this system between the two species. Streptoneury is exhibited, as in all prosobranchs (Hyman, 1967), although this is not immediately apparent as the connectives between the pleural and corresponding intestinal ganglia are so shortened as to be superficially invisible (see Fig. 16). The cerebral ganglia are indistinguishably fused to the pleural ganglia and are widely separated by a long, broad commissure overlying the gut. The buccal ganglia are distinct from the cerebral ganglia but again the connectives joining them are very abbreviated. A well developed pair of connectives, however, links the cerebral and pedal ganglia. The small buccal ganglia are fused with the cerebral ganglia, an unusual condition found only in some pteropods and neogastropods (Bullock & Horridge, 1965). A single commissure, ventral to the gut, connects the buccal ganglia. The supra- and sub-intestinal ganglia, sometimes referred to as visceral, mantle or parietal ganglia, are asymmetrically placed. The primary connective of the suprainterstinal ganglion comes from the right pleural, while the sub-intestinal is connected primarily through tracts coming from the left pleural. The visceral ganglion is accompanied by two accessory ganglia. The cerebral ganglia give rise anteriorly to three pairs of nerves whose main field of innervation is the body wall and muscles along the length of the proboscis, and the lip region. A pair of large tentacular nerves arises more laterally, supplying not only the

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cephalic tentacles but also the region of the head lying between them. Behind these arises a pair of nerves which divide before supplying the head region posterior to the tentacles, and the mantle. The left pleural ganglion gives off a large siphonal nerve which passes into an accessory ganglion before dividing into four branches—two thick inner branches and two fine outer ones—which enter the siphon at its base. A nerve from the right pleural ganglion innervates the wall of the body cavity laterally and especially the proboscis retractor muscles. Nerves from the buccal ganglia innervate the pharynx and oesophagus as well as the anterior lip region of the proboscis. These are variously called oesophageal, pharyngeal, odontophoral or oral nerves (Bullock & Horridge, 1965; Hyman, 1967). There appear to be three pairs of these nerves in B. digitalis (Donnelly, unpubl.) but five pairs in B. rhodostoma (Read, unpubl.). Urosalpinx has no less than seven pairs (Carriker, 1943). A pair of nerves from the buccal ganglia also supplies the salivary glands. The supra-intestinal ganglion gives rise to a visceral nerve which divides shortly after its origin. The right branch forms part of the visceral loop and passes through the body wall into the mantle, where it joins the left accessory ganglion. This ganglion is joined to the visceral ganglion, which in turn connects with the right accessory ganglion. The loop is completed by a nerve tract connecting the right accessory to the sub-intestinal ganglion. The left branch of the visceral nerve innervates the posterior part of the osphradium, entering the ganglion which forms the axis of that organ (Newell & Brown, 1977), with a smaller branch supplying the posterior region of the gill. A second nerve also leaves the supra-intestinal ganglion, branches supplying the anterior parts of the osphradium and gill, while a fine branch joins the siphonal accessory ganglion. The sub-intestinal ganglion gives rise to four nerves; the nerve which forms part of the visceral loop, a fine nerve supplying the floor of the body cavity on the right side, a nerve which innervates the lateral body wall, and the columellar nerve. In most prosobranchs, the last originates from the pleural ganglia (Bullock & Horridge, 1965) but both Donnelly and Read agree that in Bullia it arises from the sub-intestinal. It runs along the floor of the body cavity before diving through thick muscle layers into the columellar muscle. Four pairs of pedal nerves leave the pedal ganglia to supply the posterior and middle regions of the foot. The anterior region is supplied by five nerves from the right pedal ganglion and four from the left. In addition, a fine pair of nerves leaves the pedal ganglia posteriorly to go to the deep regions of the floor of the body cavity, below the gut.

SANDY-BEACH WHELKS AND POLLUTION Sandy-beach whelks do not make ideal organisms for marine pollution monitoring, being both mobile and of a scavenging habit. The difficulties of breeding them in the laboratory, their restricted breeding season and the impossibility of finding fertilized eggs on demand, preclude long-term studies involving several generations and render the routine testing of gametes or developing embryos out of the question. Nevertheless, on many South African sandy beaches Bullia is the most immediately obvious member of the ecosystem and its relative abundance is fairly easily assessed without recourse to

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digging. Moreover, sandy beaches, being commonest in shallow bays and in the mouths of estuaries, are rather more prone to pollution than are rocky shores. There is thus some practical value in knowing the responses of sandy-beach whelks to polluted conditions and the concentrations of pollutants which will invoke those responses, quite apart from the light that may be thrown on invertebrate physiological responses to pollution. Changes in behaviour and oxygen consumption in Bullia, consequent on the presence of a number of common pollutants, have been studied at the University of Cape Town. Brown & Currie (1973) assessed the effects of ammonium nitrate pollution on the behaviour and respiration of B.digitalis, this being part of a project connected with the toxicity of the effluent from a local fertilizer factory. Brown, Baissac & Leon (1974) investigated responses to oil pollution both in the field and in the laboratory, while D.Muir (unpubl.) considered the effects of oil dispersants. Cuthbert, Brown & Orren (1976a,b) studied both the responses of the whelks to cadmium pollution (Fig. 17) and the uptake and accumulation of this metal by B. digitalis. Changes in oxygen consumption associated with cadmium and zinc have also been considered (Brown, Davies & Young, 1982). Phenol was the pollutant studied by Golombick & Brown (1980) and Brown (1964c) described the reactions of the whelks to hydrogen sulphide. Brown (unpubl.) has tested the toxicity of a number of factory effluents using Bullia and has also made preliminary or cursory observations on the effects of a number of other pollutants in the laboratory, including mercury, lead, and copper.

Fig. 17.—LC50 values for Bullia digitalis adults subjected to cadmium chloride pollution at 15°C and a salinity of 34.5‰: 60 h LC50=4.00 mg·1−1 cadmium; 72 h LC50=3·58 mg·1−1; 96 h LC50=0.90 mg·1−1; (from K.C.Cuthbert, unpubl.).

BEHAVIOURAL AND PATHOLOGICAL EFFECTS OF POLLUTANTS

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With the exception of detergents and dispersants, all the pollutants tested induced the same pattern of behavioural and pathological changes in Bullia. The first sign of stress is a failure of the animal to be attracted to its food. This is followed by a cessation of the burrowing response, the animal lying on the back of its shell on the sand, twisting its foot from side to side. Under field conditions, this behaviour favours transport by waves and currents, possibly to a less polluted site. During more extreme pollution stress the foot becomes discoloured, changing from its normal creamy-white or yellow colour to grey, and the foot begins to become creased and curled at its edges as turgour is lost. This creasing gradually spreads inwards towards the middle of the foot. At the same time the animal becomes increasingly lethargic, foot waving subsides and the animal is unable to retract fully into its shell. The final phase is almost complete paralysis, only small easilyoverlooked movements of the foot being apparent and the animal ceasing to respond to mechanical stimulation. Animals in which the burrowing response has ceased have always been found to recover when placed in unpolluted sea water, as do most of the whelks displaying early lethargy and slight creasing of the foot. The later stages, however, are irreversible and paralysed animals have never been known to recover. Some of the results of the above work, both published and unpublished, are shown in Table III. The most appropriate, and also the easiest, stage of stress to assess is the cessation of the burrowing response. The animals always burrow after handling if the water is not polluted, so that cessation of the response is immediately apparent both in the field and in the laboratory. The onset of lethargy, foot-creasing or paralysis is more difficult to determine

TABLE III Short-term effects of some pollutants on adult Bullia digitalis at 15° C: all metals introduced in the form of the chloride

Pollutant

Cadmium

Highest Concentration at concentration which burrowing without effect, ceases, mg·1−1 mg·1−1 0.1

0.5

Concentration causing irreversible stress, mg·1−1 0.7–0.8

96-h LC50 mg·1−1 0.9

Effect on oxygen uptake 30 % increase at 0.5 mg·1−1 20 % decrease at 0.75 mg·1−1

Zinc

1

2

3

3

40 % decrease at 2mg·1−1

Lead

0.5

1

—

—

Slight but

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significant decrease at 5 mg·1−1 Copper

0.1

0.2

0.35

0.5

70 % decrease at 0.3mg·1−1

Mercury

0.5

2

7

—

20 % decrease at 5 mg·1−1

Selenium

1

4

7

—

15% decrease at 5 mg·1−1

Phenol

10

50

1000

—

100% increase at 50 mg·1−1 80 % decrease at 1g·1−1

Ammonium 50 nitrate

60

550 for 15 h

≈300

No effect up to 1g·1−1

precisely, as is the failure of the olfactory response which normally leads to emergence from the sand in the presence of food. Even the exact time of death is more difficult to determine than is the cessation of burrowing. Such sub-lethal effects in any case provide more appropriate criteria than any LC50 (Brown, 1976). An animal suffering advanced stages of stress in the field must be considered a loss to the population long before death actually ensues, seeing that it will not recover. In the case of detergents, including crude oil dispersants, emulsifiers, and household detergents, the animals respond by retracting into their shells, where they remain until dead (D.Muir, unpubl.). Oxygen uptake ceases on retraction (Brown, unpubl.). EFFECTS OF OIL Virtually all the work referred to above was conducted in the laboratory, even where factory effluents were involved. An exception is provided in the case of pollution with crude oil, where field and laboratory observations have complemented one another. Day, Cook, Zoutendyk & Simons (1971) noted the disappearance of Bullia from the sandy beach at Die Mond, some 15 km east of Cape Agulhas, following spillage of oil from the WAFRA. As this beach was not directly contaminated by the oil, the authors suggested that the whelks might have been poisoned by eating oiled Physalia or other carrion. Brown et al. (1974) showed, however, that Bullia generally refuses food soaked in crude oil, while the few that will actually sample the food appear to suffer no short-term illeffects. These authors ascribe the death of Bullia at Die Mond to the indiscriminate use of

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three different types of oil dispersant following the spill. These dispersants were far more toxic than the oil itself (D.Muir, pers. comm.). Brown et al. (1974) also studied the effects of pollution from an oil slick of unknown origin on the Bullia populations of False Bay, recording the same sequence of reactions as those noted in the laboratory. On the two days following the spill municipal workers “cleaned” the beach by the simple expedient of burying the oil where it lay, thus ensuring that the beach would remain polluted for a long time. On these two days, many B.digitalis and B. rhodostoma appeared to be behaving quite normally and even feeding individuals were observed, while others were lethargic, refused food and did not burrow after handling. For some three months afterwards, stressed individuals were found, the larger whelks appearing to be more affected than the smaller. An occasional individual had entered the paralytic phase of stress. In the laboratory it was found that actual contact with crude oil was a critical factor in determining its toxicity. Contact with even very small amounts of crude oil rapidly led to death, even when the oil had been weathered for some days, and post-mortem examination revealed patches of oil on the gill and osphradium as well as on the foot. In contrast, the whelks were very resistant to the soluble fractions of the crude oil tested, at least as far as short-term effects were concerned. These results indicate that an oil slick washing ashore during high tide may have less immediate effect on Bullia populations than similar pollution at low tide and that B. rhodostoma, frequently occurring higher up the shore, is more vulnerable to oil pollution than is B. digitalis, which in turn is more vulnerable than the infratidal B. laevissima. EFFECTS OF POLLUTION ON OXYGEN UPTAKE In general, pollutants lower the metabolic rate, as evidenced by a decrease in oxygen consumption. This applies to zinc (Brown et al., 1982), lead, mercury, copper, selenium and crude oil (Brown, unpubl.), although ammonium nitrate was shown to have no effect on respiration up to quite high concentrations (Brown & Currie, 1973). Two pollutants, cadmium chloride and phenol, raise the respiratory rate when present in low concentrations but reduce it in higher concentrations (Brown et al., 1982; Golombick & Brown, 1980). In both cases the elevation of the rate of oxygen uptake at low concentrations may be due to an uncoupling of oxidative phosphorylation. It is interesting that changes in salinity have the same effect as the majority of pollutants in lowering the respiratory rate (Brown & Meredith, 1981). The part played by reduction or cessation of the respiratory current in this phenomenon, possibly followed by partial anaerobiosis, is not yet clear. The results on Bullia compare quite closely with those of MacInnes & Thurberg (1973), who assessed the effects of various metals on the behaviour and oxygen consumption of Nassarius obsoletus, a species whose responses to pollution have much in common with those of Bullia. As in the case of Bullia, metallic ions were found to reduce the rate of oxygen uptake in Nassarius, the exception being cadmium, which increased it at low concentrations. For the most part the levels of pollution causing sub-lethal stress or 50% mortality in Bullia appear to be unexceptional and, while far higher than those causing responses in

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embryonic and larval forms (Brown, 1974, 1976; Brown & Greenwood, 1978), compare well with those affecting aquatic adult forms of both vertebrates and invertebrates. For example, the no-effect level of phenol on B. digitalis is about 10 mg·1−1, which agrees surprisingly well with the accepted toxicity of phenol to fish (D.A.Lord, pers. comm.) as well as to a variety of marine invertebrates (Smith, 1974). ACCUMULATION OF METALS Little work has been published on the possible accumulation of metals and other toxins by members of the Nassariidae. Cuthbert et al. (1976a) reported, however, very high concentrations of cadmium in the Ou Skip population of B. digitalis, sampled in midwinter, 1975. Whole Bullia had more than 106 times the concentration of cadmium in sea water sampled from the same site at the same time. The cadmium content of the viscera proved significantly higher than that of the whole body by a factor of nearly 7, reaching values over 350 mg·1−1 wet weight. These are among the highest values recorded for marine invertebrates. K.C.Davies (in prep.) has assayed the concentrations of both cadmium and zinc in other Bullia populations on the western and southern coasts of South Africa. She finds that B. digitalis on the western coast have higher levels of cadmium and zinc than southern coast populations of either B. digitalis or B. rhodostoma. Bullia appears to be an efficient accumulator of both metals but there is no evidence that the levels of the metals increase with increasing size of adult animal. ADENYLATE ENERGY CHARGE The

adenylate

energy

charge,

as

given

by

the

formula

where ATP, ADP and AMP are the concentrations of adenosine 5′-triphosphate, -diphosphate, and -monophosphate, respectively, has frequently been used in recent years to assess the physiological state, and hence pollution stress, in a variety of organisms, including gastropod molluscs (Ivanovici, 1977, 1980; Rainer, Ivanovici & Wadley, 1979). Brown & Gedye (in press) have measured the adenylate energy charge in B. digitalis, both in the field and under various conditions of stress in the laboratory. They found that the AEC of natural populations, under polluted conditions, varied from just below 1 to 0.5. This latter figure is very close to what is generally considered to be the lower limit for physiological wellbeing. In the laboratory, values for the AEC were maintained if originally high or were increased from low to high values, under a variety of conditions including burial in sand under stagnant water and enforced activity in strongly stirred and aerated water in the absence of a soft substratum. Even under totally anoxic conditions an AEC of 0.77 was maintained during a 4-day experiment. Pollution in the form of 1.1 mg·l−1 cadmium chloride also

had little effect on AEC values. Brown & Gedye conclude that, while measurements of ATP concentrations and AEC values are of considerable physiological interest, they are too insensitive to stress and too dependent on other, largely unknown, factors, to be of any great value in the

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assessment of pollution in the sea. PARASITOLOGY AND PATHOLOGY PARASITES The presence of platyhelminth parasites in Bullia, and particularly the association of larvae of digenetic trematodes with the digestive gland and gonads, has been noted by Anantaraman (1963), Brown (1971b), and Reimer (1975, 1976). The numbers of these parasites found in the South African species are usually small but tend to increase in captive whelks and damage to the internal organs may occur. Small nematode worms have occasionally been found in the mid-gut of Bullia and one individual of B. rhodostoma had a number of small cysts in the pericardium (Brown, 1971b). No ectoparasites have ever been observed, although algae and hydroids may grow on the shells of those subtidal species which tend not to bury themselves completely in the sand. W.Liltved, of the South African Museum, has supplied the author with specimens of B. laevissima dredged in Algoa Bay, each of which had numerous egg cases of a marginellid gastropod attached to the shell. ELIMINATION OF FOREIGN PARTICLES Although Bullia does not appear to remove or immobilize its trematode parasites, other invaders, including bacteria, may be dealt with efficiently. Brown & Brown (1965), following the methods of Brown (1964a), injected thorium dioxide in the form of “Thorotrast” into the pedal sinuses of B. laevissima and B. digitalis, and observed the course and accumulation of the particles in various tissues by means of radiographs. Some animals were killed during the experiment so as to X-ray various organs separately, and wax-embedded sections were viewed by combined dark-ground and phase contrast illumination under oil immersion to show up individual thorium particles. Haemolymph was extracted at intervals from living whelks in order to study the uptake of thorium by the haemocytes and the effects of the particles on haemocyte numbers. Bullia was found to be capable of eliminating such foreign particles from its body efficiently. That they are phagocytosed by haemocytes, which then migrate out of the body, is in keeping with previous work on the Mollusca. Both radiographs and sections showed that several routes may be followed by laden haemocytes in reaching the outside of the animal. The chief of these involves the heart, the pericardial cavity, the lumen of the kidney and the mantle cavity, in that sequence. It is logical to assume that the renopericardial canal and the nephropore are also involved, although such involvement could not be observed directly in the experiments. The chief pathway for migrating, laden haemocytes would thus appear to be from the haemolymph, through the wall of the heart (both auricle and ventricle) into the pericardial cavity, then through the renopericardial canal into the lumen of the kidney and out through the nephropore into the mantle cavity, from which the cells will be scoured by the respiratory current. The migration of haemocytes through the spongy heart wall is not as laborious as might be supposed as

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there is no endothelium and the muscles are everywhere invaded by blood spaces. Thus haemocytes may penetrate to the heart’s epithelium without facing obstruction. An excretory filtration of the haemolymph may also occur across this epithelium in prosobranchs (Potts, 1967). Minor routes for the migration of haemocytes are through the pericardial wall from the surrounding tissue, through the tissues of the kidney into its lumen, and through the mantle wall from the pallial blood spaces.

THE HAEMOCYTES Haemocyte counts of control animals were found to be variable, not only between different animals but between different sites in the same individual. In general, the haemolymph drawn from the heart and arteries gave higher counts than fluid from the veins and sinuses. During the experiments recorded by Brown & Brown (1965), all samples for haemocyte counts were drawn from the buccal sinus. This greatly reduced the variability, particularly in whelks kept for some days in the laboratory. Two types of haemocyte were clearly distinguished, although intermediate forms were found. These were lymphoid cells and granular macrophages, the latter being more numerous and larger than the former. Mitotic divisions appeared to be confined to the smaller lymphocytes. In samples from freshly-collected whelks, mitotic figures were observed less frequently than once in 2000 cells (total cell count), but the mitotic index increased after injection of thorium dioxide; this increase was most marked in the first 3 days after injection, indices of 11 to 14 being common by the fourth day. The index continued to increase, but more slowly, until maxima of 16 to 18 were reached after about 7 days. The number of mitotic figures then began to decrease slowly but indices of 4 to 7 were still apparent 10 weeks after injection of the particles. The total cell count began to increase only about 4 days after the injection, counts during this period being either stable or showing a slight decrease. By the sixth day, however, counts of >20 000 cells·m−3 of haemolymph were recorded. Only the macrophages were found to ingest thorium particles, although the majority withdrawn in the haemolymph were found to be free of such particles at all stages of the experiment. This confirms the observation made from sections that laden macrophages tend to stick to arterial and-other surfaces. It is probable that the macrophages originate from lymphocytes.

GENETICS Miss F.M.Da Silva (in prep.) has begun a karyotype analysis of a range of Bullia species from South Africa and has so far completed preliminary investigations on females of B. rhodostoma, B. pura, and B. digitalis from Cape Peninsula beaches (Fig. 18). The techniques employed were modified after Stern (1975), the tissues examined being ovary, hepatopancreas, and gill. In all three species the diploid (2n) number of chromosomes is 64. This is relatively high, the diploid number reported for gastropods in the literature ranging from 10 to 88 (Murray, 1975).

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This finding is of interest in view of the suggestion of Butot & Kiauta (1969) that the primitive members of molluscan groups tend to have large numbers of chromosomes, the numbers being reduced in more advanced forms. The results of Da Silva suggest otherwise or at least that there are exceptions, as Bullia is an advanced and specialized member of the most advanced superfamily of Neogastropoda. The results may, on the other hand, support the contention of Buch (1965) that molluscan chromosome numbers tend to increase with increasing specialization. Female B. digitalis collected in False Bay (Muizenberg) showed a tendency to polyploidy in both hepatopancreas and gill. Brown (1971b) was struck by the variety of colour patterns to be found in Bullia, and particularly in the B. digitalis population on Muizenberg Beach. McGwynne (1980) has paid further attention to this polymorphism in the B. digitalis populations on Maitland River Beach and King’s Beach. She distinguishes six colour varieties: cream with purple rays, cream with brown rays, purple, orange, cream, and grey. There is some evidence from the frequency distribution of these patterns that they may be genetically determined, although the possibility of diet playing a rôle cannot be dismissed. Brown & Da Silva (1979) discovered a genetic difference between females of the B. digitalis populations from the west and east coasts of the Cape Peninsula (Ou Skip and Muizenberg), with respect to oxygen uptake at low temperatures. Genetic isolation of these two populations is being further investigated using electrophoretic techniques.

CONCLUDING REMARKS Bullia has so far proved a most rewarding animal to study. It is easy to collect throughout the year and to maintain for long periods in the laboratory. It is robust, responding well to various experimental techniques and recovering rapidly from surgery or the extraction of body fluids. Moreover, the latter procedure is far easier than in the vast majority of molluscs. The animal’s relative reluctance to withdraw into its shell also facilitates various types of experimentation. The whelk is a convenient size, being large enough for a wide variety of techniques to be adopted, yet requiring relatively little space in the aquarium or constant temperature room. It requires neither running sea water nor carefully controlled conditions of temperature or oxygen tension. Indeed all that prevents it from being an ideal experimental marine invertebrate is its restricted breeding season and the fact that it will not at present breed in the laboratory. In view of its excellence as a research animal, it is perhaps not surprising that its investigation has thrown light not only on the anatomy, ecology, physiology, and pathology of the genus Bullia itself, and on the family Nassariidae, but has had far wider implications for an understanding of the Prosobranchiata and indeed of the Mollusca as a whole. Valuable insight has

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Fig. 18.—Sections of Bullia hepatopancreas fixed in acetic acid and stained with Giemse stain to show the chromosomes: A, Bullia rhodostoma; B, Bullia digitalis; (photographs by F.M.Da Silva, previously unpubl.).

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also been gained into the lives of sandy-beach animals in general and of carnivorous, scavenging psammophiles in particular. The life of Bullia is clearly dominated by its restricted and erratic food supply, so that its behaviour and physiology show many adaptations to this condition, some being concerned with finding and ingesting a wide variety of food efficiently, others with energy conservation. Undoubtedly further studies, including metabolic investigations, will demonstrate additional energy-conserving features. This may well be true, for example, of the rôle played by anaerobiosis, which has been shown in several other molluscs to be not necessarily the wasteful process once believed but a method which possibly allows the animal to decrease its metabolic expenditure to a marked degree during periods of inactivity. Assimilation efficiency and overall production efficiency remain to be studied but high values are predicted. It is also probable that we are on the brink of discovering the biochemical mechanism behind the essentially flat rate-temperature curve, a further energy-conserving device which is characteristic of B. digitalis at all levels of activity and is found in many other molluscs during relative inactivity or quiescence. Indeed it is the biochemistry determining many of Bullia’s now well-established physiological features which still needs to be investigated in greater detail. This is not to say, however, that biophysical phenomena associated with the animal are without interest. On the contrary, the light thrown on gastropod locomotory activities and fluid pressures has been considerable and the muscle-on-muscle system discovered in Bullia columellar muscle, allowing that muscle to push, in addition to exerting a force when it shortens, now appears to be a feature of prosobranchs in general and perhaps of all gastropods. This too requires further investigation. This has not been, therefore, a review of a completed field of study but rather a progress report on an animal which may be expected to yield further interesting and significant results in the future. It has been written with such expectations in mind and, in the ultimate analysis, in praise of Bullia itself.

ACKNOWLEDGEMENTS Much of the work reviewed or reported for the first time in these pages was made possible by the enthusiasm and dedication of students and research assistants working under my direction. They have included at various times Misses F.M.Da Silva, D.Donnelly, F.L.Meredith, A.Gedye, L.J. Beekman, D.Banks, A.B.Currie, Mrs K.C.Davies, Mrs T.Golombick, and Messrs R.Bally, P.Du Preez, G.Read, and J.C.Allen. I thank Drs A. D.Ansell and A.McLachlan for their critical reading of the manuscript.

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RECENT STUDIES ON THE BIOLOGY OF INTERTIDAL FISHES R.N.GIBSON Dunstaffnage Marine Research Laboratory, Oban, Argyll, Scotland

Oceunogr. Mar. Biol. Ann. Rev., 1982, 20, 363–414 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION The biology of intertidal animals is probably better known than any other group of marine organisms. There are numerous reasons why this should be so, but paramount must be the intrinsic interest in the ways such animals adapt to the fluctuating conditions in which they live. Equally important from a practical viewpoint is that intertidal animals are usually obtainable in large numbers, require no complex equipment for their capture, and because of their general hardiness, make good experimental subjects. Consequently, the literature covering the biology, in its widest sense, of intertidal species is very extensive. Many attempts have been made to summarize this vast body of knowledge ranging from numerous popular accounts through the classic works by Ricketts & Calvin (1939) and Yonge (1949) to general texts (e.g., Carefoot, 1977) and detailed reviews (e.g., Newell, 1979). Common to nearly all these works, however, is that intertidal fishes receive little, if any, attention. In recent years there has been an increasing interest in this group and the available information up to 1969 was summarized in an earlier paper (Gibson, 1969a). Since then other works have dealt specifically (e.g., Moring, 1979b) or partially (e.g., Fitch & Lavenburg, 1975; Horn, 1980), with intertidal fishes. In the past decade this interest has continued to grow and the emphasis has shifted perceptibly from an initial, essentially descriptive, phase towards a more experimental and analytical approach. The geographical area covered by such studies has also been greatly extended although they still tend to be concentrated in Europe and on the Pacific coast of North America. Observations on African, Australasian, and South American species are for the most part sadly lacking. The present review is intended to be an updated version of the earlier one (Gibson, 1969a). It covers the literature from 1969 onwards although, where particularly relevant to discussions of recent work, papers quoted in the first version have been included together with those overlooked at that time. A particular effort has been made to include unpublished theses which contain much valuable information.

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DEFINITION AND CLASSIFICATIONS The term “intertidal fishes” as opposed to “littoral” or “shore fishes” has been used in the title to avoid ambiguity as much as possible. “Littoral” is used in some works (e.g., Klausewitz, 1972) to include subtidal regions and “shore fishes” are frequently referred to as those found close to the shore with variable outer depth limits (e.g., Hobson, 1968; Derickson & Price, 1973; Lubbock, 1980). The decision as to which species to include in a review such as this is somewhat arbitrary but the definition of an intertidal fish used here is “a species which relies on utilization of the intertidal zone for completion of all or an essential part of its life history”. Such a definition takes into account the much greater mobility of fishes when compared with other intertidal animals and does not exclude those species which inhabit the intertidal zone only at certain states of the tide or seasons of the year. Very few species spend all their life intertidally because most have a planktonic larval phase but many, after metamorphosis, may remain in the intertidal zone until they die. At the other extreme some are found stranded in pools on occasions and must be regarded as accidental and very temporary inhabitants. Several attempts have been made to classify intertidal fishes, most of which are based on those found in intertidal rock-pools. Starting with Breder’s (1948) classification of Bahamian tide-pool fishes as typical, casual or accidental inhabitants, Gibson (1969a) expanded these categories into true and partial residents and tidal and seasonal visitors to cover all types of shore. Thomson & Lehner (1976) studying tide-pools in the Gulf of California defined a species’ status as being either a primary or secondary resident or as a transient in an attempt to incorporate relative mobility into the definition. Moring (1979b) preferred the terms seasonals and year-round residents for American Pacific coast species. Potts (1980) retained the separation between resident and transient species but subdivided the latter into accidental, tidal and seasonal visitors. The differences between the various classifications are relatively minor, however, and all authors agree on the distinction between residents and transients/visitors (see also Vivien, 1973; Arruda, 1979a). Generally speaking the relative proportions of residents and transients of any intertidal fish assemblage will depend to a large extent on the amount of cover available. On rocky shores where pools, crevices, boulders, and clumps of algae provide large amounts of cover, the resident component will be high. On sandy or muddy shores, on the other hand, cover may be non-existent and the great majority of fish will leave such shores on the ebbing tide unless they become stranded in intertidal pools or are capable of burying themselves in the substratum (as in the weaver fish—Trachinidae—for example, see Lewis, 1976), associating with burrowing invertebrates (e.g., Clevelandia ios, MacDonald, 1975; Brothers, 1975), constructing their own intertidal burrows (e.g., the goby Ilypnus gilberti, Brothers, 1975), and/or withstanding exposure to air as in the Periophthalmidae. This review will concentrate mainly on the intertidal residents and their adaptations to the particular physical and biological conditions which life in such a habitat imposes.

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Compared with the multitude of surveys of the flora and invertebrate fauna of the intertidal zone conducted in different parts of the world, those dealing with fishes are very limited. They range in detail from general surveys where fishes are mentioned incidentally (Batzli, 1969) or are included in the total fauna list, as in Penrith & Kensley’s (1970a,b) survey of shores in S.W. Africa, to those specifically designed to investigate intertidal fishes for a variety of purposes. The habitats surveyed range from mangrove swamps, which have received relatively little attention (Austin, 1971, in Puerto Rico; Berry, 1972, in Western Malaysia) although the general characteristics of these areas in the Indo-West-Pacific have been reviewed by MacNae (1968), through sheltered bays and intertidal creeks (Fox & Mock, 1968, in Louisiana; Derickson & Price, 1973, in Delaware; MacDonald, 1975 and Horn, 1979, in California; Cain & Dean, 1976, in S.Carolina), to open coasts with different degrees of exposure and geological characteristics. The fish populations of sandy or mixed sediment beaches, most of which consist of “tidal visitors” have been enumerated in Puget Sound by Kendall (1966) and Armstrong, Staude, Thorn & Chew, (1976), in S.Carolina by Anderson et al., (1977), and in Great Britain by Gibson (1973a), Lockwood (1974), Jones & Clare (1977), Kislalioglu & Gibson (1977) and Riley, Symonds & Woolner, 1979), and by Fonds (1973) on the Dutch coast. Ramachandran Nair, Luther & Adolph, (1965) studied the fish in temporary pools on the coast of South India. Most interest, however, has been directed towards rocky shores and there is a steadily increasing body of information which deals with tidepool fishes in many areas of the world. In Europe, for example, Alted, Sabater, Fernandez & Fernandez, (1974) have described the “intertidal” ichthyofauna of Minorca in the Mediterranean, Arruda (1979a,b) that on the Portuguese coast and the Azores, and Gibson (1972) on the Atlantic coast of France. Wheeler (1970) has surveyed the intertidal fishes of Guernsey (Channel Islands) and Hussain (1979) those of Great Britain in general. A series of papers by Chang, Lee & Wang (1969), Chang, Lee, Lee & Chen (1973), Chang, Lee & Wu (1977), and Lee (1980a,b) comprehensively covers the island of Taiwan, and Hayashi & Itoh (1978) describe some rock-pool gobies of Japan. Pinchuk (1976a,b) surveyed the intertidal fishes of the Kuril and Komandorskie Islands off the Kamchatka peninsula and Peden & Wilson (1976) included intertidal fishes in their extensive survey of the shallow water ichthyofauna of northern British Columbia and southeastern Alaska. Moring (1972) gives a list of intertidal fishes found in northern California and reviews earlier work in that area, and Thomson & Lehner (1976), working further south, conducted a 7-year census of the populations of two large tide-pools in the northern Gulf of California. More recently, Thomson, Findley & Kerstitch (1979) have described the intertidal and reef fishes of this area in detail. Marsh, Crowe & Siegfried, (1978) investigated the abundance of the commonly occurring intertidal clinids on the South African coast, Lubbock (1980) mentions intertidal species in his paper on Ascension Island “shore fishes” and Potts (1980) describes the results of a brief survey of rock-pools on Little Cayman Island, West Indies. Finally, although coral reefs lie outside the scope of this review, Vivien (1973) gives an account of collections made on the intertidal reef flats off Madagascar. For a recent review of the ecology of coral-reef fishes the reader should consult Sale (1980). As a digression here it may be relevant to consider the methods of collecting intertidal fishes. On sandy or muddy beaches traditional seines or small trawls are generally used,

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but on rocky shores such methods are of very limited use. Collecting by hand is as efficient a method as any on boulder shores but where rock-pools are to be sampled traps (Green, 1971b) or dip nets (Pinchuk, 1976a; Hayashi & Itoh, 1978) can be employed, although they are unlikely to catch all fish present if a complete census is required. Consequently, some form of poison or anaesthetic is often added to pools before collection begins. Rotenone or rotenone-based compounds are frequently used (Chang et al., 1969; Green, 1971a; Vivien, 1973; Thomson & Lehner, 1976) but together with other poisons such as sodium cyanide (Chang et al., 1977; Lee, 1980a,b) or Antimycin A (Cain & Dean, 1976) they suffer from the disadvantage that their effects are irreversible. The use of the anaesthetic quinaldine in the field was first described by Gibson (1967a) and has subsequently been used widely (e.g., Moring, 1970; Gibson, 1972; Davis, 1977; Marsh et al., 1978; Sale & Dybdahl, 1978; Grossman, 1979a; Potts, 1980). Its advantages are its potency, speed of action, relative cheapness and the fact that anaesthetized fish apparently recover rapidly. Dixon & Milton (1978) and Milton & Dixon (1980) have shown, however, that quinaldine markedly affects oxygen uptake and osmoregulation in Blennius pholis and that physiological recovery may take much longer than behavioural criteria indicate. Such findings suggest that inferences as to the “normal behaviour” of fish immediately after anaesthesia in the field during, say, capture-recapture experiments should be tempered with caution in future. The surveys outlined above were conducted for a single or a variety of purposes. In some cases the object was to provide a check list of species in the area (Austin, 1971; Moring, 1972; Alted et al., 1974; Armstrong, et al., 1976) often with a discussion of their zoogeographical status (Peden & Wilson, 1976; Potts, 1980). This aspect is also emphasized by Thomson & Lehner (1976), Lubbock (1980), and by Pinchuk (1976a,b) who notes that the intertidal ichthyofauna of the Kuril Islands constitutes only 6% of the species of the coastal fish fauna of the whole Sea of Okhotsk. No comparable figures have been published for other geographical areas, but on a smaller scale Anderson et al., (1977) consider that the fauna of tidal pools on South Carolina beaches is a depauperate derivative of that in the surf zone. Thomson & Lehner (1976) calculated that the primary residents constitute 48% of the species, 42% of the individuals but only 13% of the biomass of the dominant fishes found in two tide-pools in the Gulf of California. In Taiwan primary residents make up only 20–30% of the total tide-pool fish fauna whether judged by species or numbers (Lee, 1980a,b). These figures probably vary from season to season, however, particularly in higher latitudes because in all surveys conducted over more than a few months a marked seasonal change in the abundance and species composition was noted. On the other hand, Thomson & Lehner (1976) found that the long term species diversity and structure of the intertidal fish community which they studied varied little over seven years, in spite of such disturbances as repeated defaunation and a winter kill caused by abnormally low sea temperatures. Notwithstanding the numerous surveys, only a few attempts have been made to estimate population sizes quantitatively. All such attempts, of which Moring’s (1976) is probably the most accurate, show, however, that densities rarely exceed a few individuals/m2 (Table I). The expression of density as numbers/unit area is obviously logical for topographically uniform shores, sandy beaches for example, but may not be so for highly irregular rocky shores. For practical reasons, collections are made in the latter

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habitat at low tide when fishes tend to be concentrated in favourable areas such as pools. Densities in these low-tide refuges may, therefore, be considerably higher than those at high tide when the fishes may disperse over areas inaccessible at low tide. This difficulty of expressing densities on rocky shores in a meaningful way is reflected in the different measures used by the authors cited in Table I. Some quote densities as numbers/unit area, others as numbers/unit volume. The latter measure may be useful for species which are known to remain in pools at high tide, the former if they are known to disperse. Clearly, resident fish densities on rocky shores must be interpreted with caution and used only in conjunction with a knowledge of the extent of movement at high tide.

TABLE I Estimates of population density of intertidal fishes

Species

Location

Shore Density type

Pomatoschistus minutus

Wadden Sea North Sea

Sand

7.4×10−4−0.25/m2 1.5–10−3−0.45/m2

Varies seasonally Fonds, 1973 and annually Fonds, 1973

Pleuronectes platessa (O-group)

North Sea

Sand

<0.1–>0.5/m2

Varies with depth, Lockwood, season and year 1974

Wadden Sea

Sand

0–0.5/m2

Varies seasonally Kuipers, 1977 and anually

Japan

Rock

3–14/m3 of pool volume

Density dependent Sasaki & on shore level Hattori, 1969

Chasmichthys dolichognathus

Comments

3.7–6.2/m3

Chasmichthys gulosus Clevelandia ios

California

Mud

3–20/m2

Varies spatially

Periophthalmus cantonensis

Korea

Mud

1.5/m2

Density fluctuates Ryu & Lee, over tidal cycle 1979 (0.5–2.6/m2)

Various species, mainly (>70%)

California

Rock

≈2/m

British Columbia

Rock

≈0.4–1/1 of pool volume

Xererpes fucorum California

Rock

≈0.01–0.3/m

2

MacDonald, 1975

Fluctuates seasonally

Moring, 1976

Density depends on exposure

Green, 1971a

Fluctuates seasonally

Burgess, 1978

Oligocottus maculosus Oligocottus maculosus

2

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Rock

4/m2 8/m2 2/m2

Density in main micro-habitat varies with depth

Various species of Gulf of clingfish California

Rock

Up to 150/m2

Low tide estimate Critchlow, vary patchy 1972 distribution

Gobiesox rhessodon

Rock

Up to 0.3–0.5/m2 of pool surface

Density varies with tidal level, mean value 0.17/m2

Lipophrys canevai L.adriaticus Adriatic L.dalmatinus

California

Zander, 1980

Wells, 1979

MORPHOLOGY A general account of the morphology of intertidal fishes was given in the earlier review (Gibson, 1969a) and will serve as a starting point for this section. Since that time a few works have referred in general terms to the adaptive features of such fishes (Marshall, 1971; Moring, 1979b) emphasizing their small size, negative buoyancy, cryptic colouration, and fin modifications which have evolved to minimize displacement by turbulence and wave action. In addition to these general accounts, there are many works which refer to morphological adaptations to the intertidal habitat. Each deals with one or a number of characters, but reference to these papers will be deferred to later sections which deal with adaptation to specific conditions such as fluctuating salinities and temperatures and the problems of resisting desiccation. The morphological characteristics of particular groups of intertidal fishes have been described with varying degrees of detail by several authors. MacNae (1968) reviewed the adaptation of mudskippers to their mangrove swamp environment and Brothers (1975) in an extensive study of the small gobies Clevelandia ios, Ilypnus gilberti, and Quietula y-cauda found significant differences in several ecologically important characters which included colouration, size of the gas bladder and mouth, gill raker spacing, and epidermal specializations associated with burrowing. Zander (1972b, 1973, 1979a) used morphological and other characteristics to outline the evolutionary patterns of blenniids in the eastern Atlantic and Mediterranean. The following rather heterogeneous collection of references contains studies which do not stress adaptation as such but nevertheless are included for the sake of completeness. Hesthagen & Koefoed (1979) have reviewed the existence of swimbladders in the Gobiidae with particular reference to the sand goby (Pomatoschistus minutus). In Norwegian populations of this species, swimbladder volume drops sharply from 4.7% in May to 2.5% in June. The reduction in volume results in a change from neutral to negative buoyancy and is considered to be an adaptation to life, and particularly reproduction, in shallow waters during the summer after having spent the winter offshore. Sierra (1974) investigated the relationship between the meristic and morphological characters of Blennius fissicornis and the salinity of its habitat on the Uruguayan coast and the La Plata river. Papaconstantinou (1977) made a comparative study of the skulls of 13 Mediterranean blennies from a systematic viewpoint but

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concluded that, because of many common morphological characters, skull anatomy was not of great value in clarifying phylogenetic relationships. Cataudella & Civitella (1975) used chromosome morphology in their study of the relationships within the genus Blennius and Arai & Shiotsuki (1974) described the chromosomes of six species of blennioid fishes. Papaconstantinou (1976) examined the morphology of the thyroid gland in numerous blennies and Tamura & Honma (1977) the histology of the thymus in various Japanese gobies. Histological studies of the eyes of several species have been made by Yew & Wu (1979) who found that the retina of Periophthalmus chrysospilos contained more rods than deeper living species and Kunz & Kelly’s (1973) investigation of the retina of the common goby (Pomatoschistus microps), which lives in very shallow water, showed that it was of a pronounced diurnal type and adapted for exposure to bright light. The related P. minutus has an irridescent cornea which probably acts as a sophisticated sunshade protecting the eye from bright down welling light (Lythgoe, 1975). Loew & Lythgoe (1978) examined the visual pigments of 18 species of fishes and were able to correlate the wavelength of maximum absorption (λmax) with the spectral quality of the water in which the fishes lived. Intertidal and immediately subtidal fishes, classified by Loew & Lythgoe as “littoral coastal species” formed a group whose λmax could be distinguished from other groups living in deeper water, on coral reefs or in fresh water (see also Lythgoe, 1979).

ADAPTATION TO THE PHYSICAL ENVIRONMENT The very nature of the intertidal zone, exposed as it is to marked fluctuations in environmental conditions, means that those fishes which live there must be behaviourally and physiologically adapted to cope with such fluctuations. The degree to which fishes are able to tolerate or resist change separates the residents from the transients in physiological terms rather than temporally or spatially as in the original definitions. To be an intertidal resident and remain above low water mark at low tide a species must be able to deal with the consequences. These consequences include exposure to changes in salinity, temperature, availability of oxygen, the risk of desiccation, and exposure to nonaquatic predators. Such stresses may be alleviated somewhat by remaining in pools, crevices or beneath stones at low tide when such refuges are available; when they are not then the alternatives are to construct refuges in the form of burrows or simply to bury in the sediment. Species which are not adapted to cope with drastic changes in environmental conditions leave the intertidal zone on the ebbing tide and are classed as tidal visitors. Fishes may also have to contend with turbulence and wave action at high tide and resident species, particularly those on rocky shores, have evolved structural modifications to resist displacement under such conditions. Again, transient species are more likely to avoid the intertidal zone when turbulent conditions prevail. It is the intention of this section to discuss the reactions of intertidal fishes to physical changes in their environment after examining the extent to which such changes occur, bearing in mind that specific factors such as temperature rarely act in isolation and that the limits of tolerance of species are best defined in terms of “multivariable response surfaces” (Newell, 1979).

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HABITAT STUDIES The physical factors affecting intertidal animals have been extensively reviewed by Newell (1979) and need not be discussed further here. Canning (1971), Daniel & Boyden (1975), Wright & Raymond (1978), Congleton (1980), and Davenport, Busschots & Cawthorne (1980) describe conditions in rock-pools and emphasize that such small bodies of water can undergo marked changes in their physical conditions over the lowtide period. All these authors, as well as Green (1971a), record, however, that pronounced vertical stratification may develop in such features as oxygen concentration, temperature, and salinity. This stratification may protect fishes in pools if they remain on or near the bottom (see Milton, 1971). In certain cases horizontal gradients may also be formed where the shore consists of long gulleys at right angles to the water line (Norris, 1963). Continuous records of one or more environmental variables on the shore are rare, but Green (1971a) and Nakamura (1976b) were able to monitor tide-pool temperatures for long periods. They demonstrated the sudden change which can occur in summer when the flooding tide enters high level pools as well as the difference in temperature extremes between high and low level pools. Mention should also be made of the long term study by de Wilde & Berghuis (1979) of cyclic temperature fluctuations on a tidal mud flat on the Dutch coast. Studies of environmental conditions in microhabitats such as among exposed clumps of algae or beneath stones are also uncommon but the results may be more ecologically meaningful than “spot checks” in pools. Barton (1978) measured intertidal microhabitat temperatures and salinities in California and found only slight fluctuations. Horn & Riegle (1981) also working in California found that such microhabitats were slightly cooler than the surrounding air in summer and slightly warmer in winter. They also point out that temperature fluctuations in California are further reduced because extreme low tides occur in the early morning in summer and in the afternoon in winter. Marliave (1981), however, determined that temperatures beneath algae high on the shore near Vancouver were noticeably lower than that of the surrounding bare rock. SALINITY Marked salinity change caused by dilution or evaporation is most likely in the intertidal zone at low tide and the extent of such change experienced by an individual fish will depend upon its position on the shore in relation to both the tidal level and the type of its refuge. As mentioned above fish in pools are unlikely to experience such drastic changes in salinity as those isolated in gulleys receiving direct freshwater run-off. Early experimental studies on the reaction of intertidal fishes to salinity change (see Gibson, 1969a) concentrated on their tolerance of both hypo- and hypersaline conditions. There have been few such studies recently but all confirm the earlier results in that intertidal fishes are generally euryhaline. Thus, Foster (1969) notes that the cottids Cottus bubalis and C. scorpius will indefinitely tolerate 10% and approximately 36% sea water respectively, the Indian mudskipper Periophthalmus dipes survives exposure to 24–47% (Bhan & Mansuri, 1978) and Ramachandran Nair et al., (1965) found several fish alive in isolated temporary tidal pools with salinities exceeding 88‰. These fish, however, were

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not true intertidal residents but better classed as “accidentals”. The estuarine and brackish water goby Pomatoschistus microps will tolerate a gradual increase in salinity up to 80‰ and a decrease to about 8‰ although some individuals can survive in distilled water (Kunz, 1969). Kunz also showed that this species which migrates into areas of higher salinity to breed does not alter its tolerance during the migratory period. The influence of salinity and its relationship to temperature on the eggs of P. microps was studied by Fonds & van Buurt (1974) who found that highest survival was obtained at 25‰ salinity and 20 °C. Comparison with two other species P. minutus and P. pictus which are more open coast and deeper water forms demonstrated, however, that salinity, although important, is probably not as vital as temperature in controlling the breeding success and hence distribution of these three species on the Dutch coast. Similar conclusions as to the relative effects of temperature and salinity on the metabolism, food consumption, and growth of P. microps were reached by Tolksdorf (1978). The distribution of the western North American Anoplarchus purpurescens and Pholis ornata, which are both good osmoregulators in from 0–75‰ salinity (Barton, 1978, 1979) may also be reflected in their osmotic tolerance because both are more euryhaline than the open coast Xiphister atropurpureus (Evans, 1967a,b,c). There is also a slight difference in the resistance to hyperosmotic stress over time of Anoplarchus purpurescens and Pholis ornata which probably allows the former species to inhabit open coasts and lower bay areas, whereas the latter is restricted to bays and estuaries (Barton, 1979). The sodium chloride and water balance of the European P. gunnellus was studied by Evans (1969) who suggested that this species cannot survive in fresh water because it lacks a mechanism for preventing loss of chloride from the body. Doyle (1974) briefly investigated the salinity tolerance of the same species and found it could only survive in one third strength sea water for 60 h or for 22 h in double strength sea water (salinity not given); his main finding, however, was that the tubular reticulum of the chloride cells in the gills is intimately involved in the regulation of ion transport. The majority of experiments on the effects of salinity involve sudden changes from one salinity to another and while this may occur naturally on a flooding tide, gradual changes are more likely on the ebb tide. One approach to this problem has been taken by Davenport & Vahl (1979) who investigated the effect of cyclic changes in salinity of 0– 34‰ on Blennius pholis using both square wave and sinusoidal salinity cycles of approximately tidal period. Under these conditions the fish were able to maintain remarkable physiological stability, the blood osmolality remained constant and changes in heart rate and opercular beat were negligible. Small but significant increases in oxygen consumption were found at low and rising salinity levels but they dropped back to “routine levels” at high and decreasing salinities. Davenport & Vahl suggest that the increase in oxygen consumption in rising salinities possibly reflected an increase in physical activity and postulated that in the wild, increasing salinity may act as a cue for the resumption of searching behaviour. Milton (1971) also studied oxygen consumption in B. pholis in hypotonic salinities and found that consumption was initially reduced but later rose again, a rise which was probably associated with acclimation, an increase in glomerular filtration rate and the increasing stress of ionic regulation in such conditions. Another interesting finding was that adults are more euryhaline than juveniles, probably because the latter have less efficient kidneys and a greater gill area/unit body weight than

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the adults. The Mediterranean blennies B. pavo and B. sphinx also adapt well to salinity changes (Müller, Boke, Martin-Neumann & Hanke, 1973) although Zander (1972a) classified the latter as stenohaline. Both species show only small changes in the osmolarity of their body fluids when exposed to lowered salinities and it was found that adrenocorticotropin (ACTH) and cortisol lower the sodium content of B. sphinx (Müller et al., 1973). Luteinizing hormone (LH) and follicle stimulating hormone (FSH) decrease the number of mucous cells in the skin of the same species (Blüm, 1972) as well as numerous others (Blüm & Fiedler, 1972) with a possible effect on their osmoregulatory abilities, although in Leptocottus armatus, skin goblet cells which increase in number in hyposmotic media are not controlled by pituitary hormones (Marshall, 1979). In an earlier study on the goby Gillichthys mirabilis Marshall (1977) found that the skin may play some active rôle in osmoregulation because evidence was obtained for active ion transport across it, and suggested that a functional chloride-excreting pump may be present in the skin of seawater-adapted fish. Other studies of the control of osmoregulatory capabilities of Gillichthys are those of Doneen (1976) who demonstrated in vitro that water and ion movements in the urinary bladder are hormonally controlled, cortisol elevating and prolactin lowering bladder permeability to water. Owens, Wigham, Doneen & Bern, (1977) in their experiments, however, found no evidence for the effects of cortisol, growth hormones or triodothyronine on the bladder, but human growth hormone did reduce water permeability, an effect which was attributed to its inherent prolactin activity. Loretz (1979) also demonstrated the rôle of prolactin in osmoregulation. Urophysial hormones are probably also implicated in the regulation of plasma ion levels in Gillichthys (Fryer, Woo, Gunther & Bern, 1978; Marshall & Bern, 1979). TEMPERATURE Temperature is of great significance in the life of intertidal fishes; its effect on the great majority of metabolic processes means that it controls their vertical, seasonal, and latitudinal distributions. The degree to which particular species can penetrate and survive in the upper intertidal regions depends to a large extent upon their ability to tolerate the extremes of temperature likely to be experienced there. Thus, the Gulf of California clingfishes Pherallodiscus funebris and Tomicodon humeralis are more tolerant of high temperatures and live higher on the shore than the related Gobiesox pinniger and Tomicodon boehlkei. All species were found to be more tolerant to cold than to heat and their intertidal distribution was better correlated with cold rather than heat tolerance. In summer, T. boehlkei was the only species found in a habitat where the maximum temperature did not exceed its upper lethal maximum as determined in continuous exposure experiments. The three other species could survive the high temperatures only because they were exposed to them for relatively short periods and the upper level T. humeralis may depend upon its high capacity for water loss as a means of thermoregulation by evaporative cooling (Eger, 1971). The tide-pool sculpin Oligocottus snyderi also has an experimentally determined critical thermal maximum which is approximately equivalent to the maximum temperatures encountered in upper pools on the coast of British Columbia. It is less capable of functioning at higher temperature

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extremes than its congener O. maculosus and in all probability its upper limits are governed mainly by temperature. O. maculosus which is found higher on the shore than O. snyderi, on the other hand, has a critical thermal maximum which is greater than observed maximum pool temperatures and temperature does not limit its vertical distribution. These conclusions were reached from field and laboratory experiments on temperature tolerance which also demonstrated that sudden temperature changes such as may be caused by the tide flooding a pool were not important for either species (Nakamura, 1976b). The upper limits of two other species found on the Pacific coast of North America, the stichaeid Anoplarchns purpurescens and the pholid Pholis ornata, are probably not governed by temperature either (Barton, 1978) and temperature was considered to be of only secondary importance in influencing the intertidal distribution of other stichaeoids in California (Horn & Riegle, 1981). Populations of Anoplarchus purpurescens which are subjected to sub-zero temperatures in Alaska are able to produce an ‘antifreeze’ in their blood in winter. This substance is not present in summer and requires a combination of warm temperatures and long photoperiods before its production ceases. The same species in warmer Californian waters cannot produce the ‘antifreeze’ even when cold acclimated (Duman & De Vries, 1974). The species has also been shown to exhibit enzyme polymorphism associated with different temperature regimes in Puget Sound (Johnson, 1971, 1977). Morris (1963) in an extensive study of the effects of temperature on the opaleye Girella nigricans discovered that different developmental stages select separate temperatures in the field resulting in a differential distrubution of size groups on the shore. Prejuveniles congregated in the upper reaches of channels where temperatures were high, transformed individuals selected temperatures of about 26°C, whereas larger fish preferred the cooler water lower on the shore. Evidence was obtained that these differences in distribution were the results of active temperature selection because observations in natural and experimental temperature gradients showed that the transformed fish selected 26°C preferentially. Experiments on prejuveniles and larger fish, however, were less conclusive due to problems associated with the behaviour of these stages in the experimental thermal gradient. Norris (1963) also concluded that temperature selection may aid in directing the inshore migration of prejuveniles and such behaviour may hold them in the intertidal zone where transformation can take place and at the same time reduce predation on this vulnerable stage. The onshore-offshore migration of the eurythermal goby Pomatoschistus minutus is probably also controlled by thermoregulatory behaviour. In spring when onshore migration takes place, experimentally selected temperatures are higher than sea temperatures, but lower in autumn when the fish move into deeper, cooler water (Hesthagen, 1979). Fonds & Veldhuis (1973) compared the oxygen consumption rates of this species with the shallower living P. microps and found that in the latter species its lower and more stable respiratory maintenance costs were probably related to its habit of living in waters where temperature fluctuations are large. Where sympatric species have similar vertical distributions their response to seasonally changing temperatures apparently depends upon their geographical origin. The blenny Hypsoblennius gentilis and the sculpin Clinocottus emails, for example, have a similar habitat and latitudinal distribution in California but C.analis, being of boreal origin, adapts to a temperate habitat by accommodating to warm temperatures. Hypsoblennius

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gentilis, conversely, is of tropical origin and has accommodated to cool temperatures. Both species compensate for seasonal temperature changes by adjusting their metabolic rates (Graham, 1970a). The European Blennius pholis also adapts its metabolic level to prolonged exposure to temperature change (Wallace, 1973), allowing it to reduce the rate of utilization of food in the summer and conversely to remain active during the winter. The slow rate of metabolic adjustment of this species suggests that adjustment to diurnal or tidal temperature changes would be negligible (Campbell & Davies, 1975). Three species of the clinid genus Gibbonsia, G. metzi, G. elegans, and G. montereyensis inhabiting the Pacific coast of North America all show reactive adjustments to temperature after acclimation. These species have roughly equal vertical distributions in the lower intertidal zone, but the species with the widest geographical range, G. metzi, is capable of withstanding a greater range of temperatures than the other two. G. elegans, on the other hand, which has the southernmost distribution but experiences the widest seasonal range of temperatures had the greatest tolerance of high temperatures and showed an anticipatory reaction to temperature change, enabling it to adjust physiologically in advance of seasonal changes (Davis, 1977). Thomson & Lehner’s (1974, 1976) observations of tide-pool fishes in the Gulf of California showed that low sea temperatures were the most critical factor limiting diversity. Laboratory experiments distinguished between the thermal tolerances of temperate and tropical components of the fauna and the “natural experiment” of an abnormally cold winter supported the laboratory results. They found no evidence of a “summer kill” or of movements to deeper water to avoid high temperatures. There was, however, a correlation between sea temperature and numbers of individuals which was attributed to the summer recruitment of juveniles. The effect of a “winter kill” on intertidal fish populations on the west coast of England has also been recorded (Jones & Clare, 1977). Some species were apparently locally exterminated, while others took several years to regain their former numbers. Only one, the goby Pomatoschistus microps, which has a life cycle of about two years, reestablished itself rapidly. The arctic-boreal species Pholis gunnellus was not affected by abnormally low temperatures. TURBULENCE AND WAVE ACTION The general morphological adaptations which enable intertidal fishes to live in turbulent water were outlined in the earlier review (Gibson, 1969a). There it was emphasized that they are generally poor swimmers and keep close to the bottom. This benthic mode of life is aided by negative buoyancy resulting from the absence or reduction of the swim bladder. Specialized attachment organs are present in some groups (gobies, clingfish, and liparids), which enable them to adhere to the substratum. Such organs are not present in blennioids which use the anal fin and lower portions of the paired fins to brace themselves against water currents. A suction mechanism equivalent to those of gobies and clingfish has been described, however, in the salariine blennies Lophalticus kirkii (Zander, 1967) and Alticus saliens (Abel, 1973). In these species an air bubble is trapped beneath the pectoral and pelvic fins which, together with the mouth, serve as suction cups. In addition, the rays of the anal, pectoral, and pelvic fins of numerous species of blennies are hook-shaped and allow them to cling to rough surfaces (Zander, 1967,

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1972a, c, 1980; Mayer, 1970; Smith, 1973). Zander (1972a, c) has also described how the degree of differentiation of these fin rays is correlated with the amount of turbulence in the habitat. Species which live in the surf zone have shorter, thicker, and more strongly recurved rays than those living in deeper less turbulent water. Those rays which come into contact with the bottom are often covered with a thick cuticle (Zander, 1972a, c) whose fine structure has been described for Blennius pholis by Whitear (1970). Eger (1971) also suggests that the thickened epidermis of certain clingfish may act as a protection against abrasion. Body form in Mediterranean blennies also seems to be related to turbulence. Those species which live in turbulent regions have rounded or depressed bodies whereas deeper living species have bodies which tend to be more compressed. Lateral line structure and development (Mayer, 1972; Zander, 1972a), however, is not related to the degree of turbulence which these fish experience (Zander, 1972a). Body size may also be related to turbulence and exposure to wave action because Critchlow (1972) found a gradient of body length with increasing exposure in different habitats in Gulf of California clingfish and other species. In general, fish were smaller on exposed coasts both within and between species than they were in more sheltered areas, partially, it was suggested, because turbulence reduces available feeding time resulting in a slower growth rate and hence smaller size. Presumably because of difficulties of observation the behaviour of fishes in turbulent waters is not well known but Abel (1973) describes how the amphibious Alticus saliens avoids breaking waves by jumping or climbing out of reach. Phillips (1977b) watched Istiblennius zebra swimming from pool to pool within wave surges or avoiding approaching waves by swimming into its shelter. Taborsky & Limberger (1980) observed that Blennius sanguinolentus greatly reduces its activity in rough conditions, but on occasions did see the fish dashed against the rocks. Direct damage or mortality caused by waves and turbulence is also difficult to estimate, although Nursall (1977) noted that Ophioblennius atlanticus frequently bore minor abrasions on the body as well as puncture wounds from sea urchins. There are also a few records of changes in abundance of species after storms. Such changes in abundance may have been caused by mortality or, more likely, by changes in the suitability of the habitat (Green, 1971a; Richkus, 1978). Differences in the ability of species to withstand turbulent conditions may partially account for the differences in depth distribution observed in the Mediterranean by Zander (1972a) and in the Pacific by Green (1971a). EMERSION Many resident species can often be found out of water, some only at low tide, other more amphibious species at all stages of the tidal cycle and it is a common finding that such fish show numerous adaptations to emersion. The degree to which such adaptations are expressed depends to a large extent upon whether the fish simply endure the period of emersion or whether they have evolved an almost terrestrial mode of life. Desiccation is one of the fundamental hazards facing animals which are basically aquatic but it is now well documented that resident species are capable of withstanding considerable water loss and that their vertical distribution on the shore (Eger, 1971; Horn & Riegle, 1981) is often related to this capability. The most comprehensive investigation

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of desiccation tolerance is that by Horn & Riegle (1981). In a comparative study of Xiphister mucosus, X. atropurpureus, Cebidichthys violaceus, Anoplarchus purpurescens, and Xererpes fucorum, they demonstrated that, depending on the species, these fish could survive out of water for 6–37 h and tolerate water losses of 16–24% of wet weight. Species having a higher initial water content were able to survive relatively longer. Survival time and water loss was dependent upon size, larger fish were able to live longer when emersed because of their greater surface area/weight ratio, a feature noted earlier for Cebidichthys violaceus by Riegle (1976). This inverse relationship between weight and the extent of water loss appears to be a common phenomenon because it has been recorded for Blennius pholis (Daniel, 1971), several clingfish (Eger, 1971), Sicyases sanguineus (Gordon, Fischer & Tarifeño, 1970), and Mnierpes macrocephalus (Graham, 1973). Actual survival times and weight loss out of water vary widely between species and are only strictly comparable on a weight specific basis at equivalent temperatures and humidities but they range from 4–6 days with a water loss of 22% in Blennius pholis (Daniel, 1971) to 2–7 h and ≈20% weight loss in Mnierpes macrocephalus (Graham, 1973). The greatest tolerable water loss recorded is the 50–60% for the clingfishes studied by Eger (1971). Intermediate values are reported for Sicyases sanguineus (Gordon et al., 1970; Marusic et al., 1981), Periophthalmus cantonensis (Gordon, Ng & Yip, 1978), and P. sobrinus (Gordon et al., 1969). The values quoted are all for high humidities in shade and still air; where comparable measurements have been made in lower humidities, sun or moving air (Gordon et al., 1969, 1970, 1978; Graham, 1973), the survival times and percentage water loss at death are much lower. Under these conditions death is probably caused more by overheating than desiccation. Gordon et al. (1970), for example, found that the body temperature of Sicyases reached 9 deg C above ambient when exposed to full sun and Graham (1973) noted that Mnierpes body temperature increased rapidly when exposed to sunlight in the absence of water. If kept moist, however, evaporative cooling kept the temperature down. Both Daniel (1971) and Eger (1971) also consider that evaporative cooling plays an important rôle in maintaining lowered body temperatures in the species they studied. Considering the marked effect of sun and wind on increasing desiccation it is not surprising that those species which are active out of water tend to avoid prolonged exposure to direct sunlight and rarely allow their body to become dry. The mudskippers frequently return to the water for brief periods before emerging again and Gordon et al., (1978) observed that Periophthalmus cantonensis was rarely out of contact with water for more than 1 minute on sunny days. Much longer periods were spent ashore at night and they can easily tolerate a full day out of water in the shade. In this species the stimulus for contact with water seems to be loss of water from the mouth and gill cavities. Ebeling, Bernal & Zuleta (1970) noted that Sicyases sanguineus is wetted by spray about once every 5 min and Graham (1973) found experimentally that Mnierpes macrocephalus can only survive out of contact with water in the sun for about 5–10 min, in good agreement with their observed emergence times. On cloudy days, however, it stays out of water for up to 30 min and is inactive for long periods out of water at night. Those species which are not as terrestrial as the mudskippers or Mnierpes but merely remain inactive when emersed at low tide exhibit behaviour patterns which tend to reduce water loss. Horn & Riegle (1981) note that stichaeids tend to associate in moist places at low tide, curl up, orientate towards the sides

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of experimental enclosures and produce a mucous film. Eger (1971) observed similar behaviour in clingfishes. These behaviour patterns and those shown by the “terrestrial” species are probably the most important means of reducing water loss because there is no evidence so far available that fish can prevent desiccation by physiological means. Daniel (1971) and Horn & Riegle (1981), for instance, found that live and dead fish lost water at comparable rates and Daniel (1971) states that Blennius pholis shows no form of physiological control over water loss or the resultant increase in body solute concentration. The physiological effects of desiccation have been studied in Sicyases sanguineus where it results in a marked decline in blood volume and increases in plasma electrolyte levels (Marusic et al., 1981) and Gordon et al. (1978) recorded differential rates of water loss in several organs of Periophthalmus cantonensis; the heart lost most water, the brain and liver the least. The only study on the resistance of the eggs of intertidal species to desiccation is that by Marliave (1981) on Clinocottus acuticeps. This species which spawns high on the shore lays its eggs beneath algae, experimental removal of the algae resulted in significant mortality in the exposed eggs. It was also shown that a 12-h exposure of eggs to dry conditions at 17°C caused complete mortality, whereas if the eggs were kept wet the mortality only amounted to about 50%. One further problem posed by being out of water is that of nitrogen excretion. The great solubility of ammonia, the main nitrogenous excretory product of teleosts, means that it is readily disposed of in water. On land, however, this is not the case and several studies have investigated the ways in which amphibious species prevent the build up of this toxic substance. Most investigations have centred on the periophthalmids. Gregory (1977) examined the excretory products of Periophthalmus expeditionium, P. gracilis, and Scartelaos histophorus when the fish were submerged and found a predominance of ammonia. Similar results were obtained by Morii, Nishikata & Tamura (1978) for Periophthalmus cantonensis and Boleophthalmus pectinirostris and they found that the nitrogen excreted by the skin was higher in urea than that from the gills. A higher ureaN/ammonia-N ratio was produced on land because the nitrogen produced originated mostly from the skin. A shift towards ureotelism out of water has been recorded in Periophthalmus sobrinus (Gordon et al., 1969), P. cantonensis (Gordon et al., 1978), and the clingfish Sicyases sanguineus (Gordon et al., 1970). In all three species the waste nitrogen is accumulated in the body and released on return to water. Later studies by Morii (1979) and Morii, Nishikata & Tamura, (1979) suggested, however, that in Periophthalmus cantonensis and Boleophthalmus pectinirostris, ammonia is not converted to urea while the fish are out of water. Morphological adaptations to a semi-terrestrial existence stem mainly from the differences in the densities of air and water. The much lower density of air means that locomotion on land requires greater muscular strength and a modification of the locomotor organs. Zander (1972c, d) showed that in the amphibious blenny, Alticus kirkii, the pectoral fin radiais are movable, the pelvis is strongly developed, the fish has short strong pelvic fin rays and the muscles are modified accordingly. There are also hooks at the ends of the anal, pectoral, and pelvic fin rays. All these features were considered to aid climbing on steep rocks while out of water. Graham (1970b) also describes similar modifications in Mnierpes macrocephalus in which the anal fin has an

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incised membrane between the rays. The ventral aspects of the body, anal, pectoral and pelvic fins are padded. Land locomotion in this species is performed by alternate flexures of the tail very similar to that described for Lophalticus (Alticus) kirkii by Zander (1967), for Alticus saliens by Abel (1973), and for Gobionellus sagittula by Todd (1976). This method differs from those used by mudskippers which include “crutching”, “skipping” and climbing on vegetation by means of the pectoral and fused pelvic fins (MacNae, 1968; Berry, 1972). The refractive index of air is also markedly different from that of water and the eyes of several amphibious species are known to be adapted for aerial vision. The eyes are frequently protuberant and movable as in Alticus kirkii (Zander, 1972c, 1974) and all mudskippers and the corneas of Dialommus fuscus (Herald & Herald, 1973), Alticus kirkii (Zander, 1974), and Mnierpes macrocephalus (Graham, 1970b; Graham & Rosenblatt, 1970) are all modified to allow better vision in air. Zander (1974) gives a detailed comparative account of eye structure in the Red Sea blennies Salarias fasciatus, Istiblennius edentulus, and Alticus kirkii. Finally, fish out of water have to contend with the problem of a change from aquatic to aerial respiration. This topic is dealt with separately in the following section.

RESPIRATION Many of the studies of respiration of intertidal fishes have been concerned with their ability to respire aerially and Graham (1976) has comprehensively reviewed the whole subject of air breathing in marine fishes. Consequently, only studies which have appeared since then need to be mentioned here. Graham records the presence of air breathing in 40 species, nearly all of which are intertidal. He notes that, in contrast to the freshwater airbreathers which mostly use the swim bladder, branchial diverticula or the alimentary canal, marine species rely mainly on their gills, skin and modified buccopharyngeal and opercular epithelia. Marusic et al. (1981) however, have recently demonstrated that the gut of Sicyases sanguineus often contains air and becomes gorged with capillaries after 24 h of emersion, suggesting that intestinal respiration may take place. Graham (1976), further concluded that in most species the metabolic and heart rates do not change greatly when the fish moves from water to air or vice versa and that in air the fish are generally proficient in the release of carbon dioxide. The ecological significance of air breathing in intertidal fishes seems to be that it enables them to exploit a habitat relatively free from competitors and predators. Their ability to remain out of water at low tide also means that they avoid the necessity of making energetically costly migrations with each tidal cycle. Studies concerned with aerial respiration which have been reported since Graham’s review have tended to confirm his conclusions. Thus, Riegle (1976) found that Cebidichthys violaceus respired at comparable rates in air and water at 10.5°C. At 14.8 and 18.0°C the aquatic rates were higher. The difference could be accounted for by the work performed by the respiratory muscles pumping water over the gills when immersed. At the highest temperature investigated (20.5°C) the aerial rate was significantly lower than the aquatic rate, an indication that the species is unable to satisfy its metabolic demands in air at this temperature. Buckley (1980) and Buckley & Horn (1980) also

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studied the respiration of Cebidichthys violaceus as well as Xiphister mucosus. They found, however, that in both species oxygen uptake in air was markedly lower (>80%) than that in water. Riegle (1976) also found that, although Cebidichthys violaceus showed an initial tachycardia on both immersion and emersion, the heart rate gradually returned to normal. In Periophthalmus cantonensis the heart rate also shows little alteration when the fish changes from one medium to another and in this species, too, oxygen uptake is similar in air and water (Gordon et al., 1978). Tamura, Morii & Yuzuriha (1976) demonstrated that both P.cantonensis and Boleophthalmus chinensis take up more oxygen from the water than from the air when free to select their habitat, but that oxygen uptake decreases when they are restricted to either medium alone. In air Periophthalmus cantonensis takes up about 75% of its oxygen through its skin whereas in Boleophthalmus chinensis oxygen uptake is shared about equally by the skin and gills. Welsch & Storch (1976) and Hughes & Datta Munshi (1979) provide detailed descriptions of the fine structure of the gills in Periophthalmus vulgaris and Boleophthalmus boddaerti, respectively. Cutaneous respiration in the less terrestrial species Pholis gunnellus, Ciliata mustela, and Blennius pholis was studied by Nonotte & Kirsch (1978). All three species are frequently found out of water at low tide. In Pholis and Ciliata, although oxygen is taken up by the skin in water, it is also consumed by it so that no net transport into the body takes place. Cutaneous respiration in air was not studied in these two species. In Blennius pholis, on the other hand, oxygen consumption by the skin was much lower so in this species there is a net inwards transcutaneous oxygen flux. In air, oxygen uptake by the skin only increased if the gills were also emersed. The oxygen affinity of the haemoglobin of this species is moderately low, but the Bohr effect is relatively high, suggesting that the fish is adapted to respire in a well oxygenated environment low in carbon dioxide. It is also fairly efficient at eliminating carbon dioxide in air (Daniel, 1971). Although the oxygen content of coastal sea water is normally high, on occasions fish isolated in pools at low tide may be subjected to lowered oxygen concentrations, particularly at night when algae in the pools are not photosynthesizing. Under such conditions certain species have been observed to emerge from pools or respire near the surface. Congleton (1980), for example, noticed that Clinocottus analis and Paraclinus integripinnis rose to the surface of pools at night and ventilated their gills in the surface water. Measurements of oxygen tensions in these pools showed that they decreased rapidly and became less than the fishes’ critical oxygen tensions. If access to the surface was denied under these conditions Clinocottus analis died within 2 h. A third species, Gibbonsia elegans was found to have a higher critical oxygen tension and a lower tolerance of anoxia than the other two species, and was not found in the pools at night. Actual emergence from pools by Clinocottus recalvus was observed by Wright & Raymond (1978). When out of water this species, which consumes oxygen in air and water at approximately the same rate, gulps air into its mouth, holds it there for a few minutes and then releases it. It does not. however, seem to be able to metabolize carbon dioxide efficiently in air. Buckley (1980) came to a similar conclusion for Cebidichthys violaceus and Xiphister mucosus. Wright & Raymond (1978) concluded that hypoxia is the probable cause of emergence of Clinocottus recalvus, although other factors may also

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be important. Similarly, Davenport & Woolmington (1981) found that Lipophrys (Blennius) pholis, Ciliata mustela, and Taurulus bubalis only emerge from water in response to low oxygen concentrations; increased temperature, carbon dioxide concentrations, alterations in pH and decreases in salinity had no effect. Eger (1971) also found that Tomicodon humeralis emerges in response to low oxygen levels rather than to high carbon dioxide concentrations. Blüm, Machemer & Bartmann (1972) demonstrated that neither the thyroid gland nor the hormone TSH are implicated in inducing Blennius pavo to leave the water. Hypoxia may also present a problem to those species which inhabit burrows. Gordon et al. (1978) found that the burrow water of Periophthalmus cantonensis was virtually anoxic and that the species has only a limited tolerance of such conditions. Congleton (1974) studying the symbiotic goby Typhlogobius californiensis, which lives in the burrows of the shrimp Callianassa where the burrow water also becomes anoxic at low tide, discovered that this fish has a very low resting oxygen consumption. It metabolizes oxygen in the gas bladder during the initial hours of asphyxia and can also respire anaerobically. Saksena & Joseph (1972) investigating the dissolved oxygen requirements of newly-hatched larvae of Chasmodes bosquianus, Gobiosoma bosci, and Gobiesox strumosus concluded that they are unlikely to suffer mortality caused by decreased oxygen concentrations at low tide. Other aspects of respiration in relation to physical factors such as temperature and salinity have been discussed in the relevant sections.

REPRODUCTION The reproductive process of resident intertidal fishes follows a fairly generalized pattern which may be summarized as follows. The male selects a spawning site in a sheltered position, usually in a hole or beneath a stone, to which he attracts one or more females. This spawning site is often situated within a territory which the male defends both before and after spawning. Batches of eggs are deposited singly on the substratum and fertilized by the male who then cares for them until they hatch. With very few exceptions recent studies have shown that species whose breeding habits were previously unknown or only imperfectly described all comply with this general scheme. Detailed experimental investigations on the factors involved in the selection of a particular spawning site do not seem to have been performed. The observation that certain species do spawn in particular places rather than at random suggests, however, that some process of selection does operate in many cases. Moosleitner (1980) for example, describes how the males of Blennius pavo selected the only available boulder for spawning in an otherwise sandy area. Omobranchus loxozonus (Dotsu & Oota, 1973) and Neoclinus bryope (Shiogaki & Dotsu, 1972b) also show evidence of site selection in that they lay their eggs in vermid shells, and gobies of the genus Pomatoschistus spawn under bivalve shells, although here there seems little evidence that a specific shell type is selected (Fonds, 1973; Vestergaard, 1976). Marliave (1975) describes an example of sitespecificity in more detail. Xiphister atropurpureus only spawns under boulders where the substratum provides interstitial spaces comparable to the cross sectional area of the body.

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A specific requirement for a particular degree of water movement is also apparent because X.atropurpureus shifts its spawning site in the course of the breeding season. Early in the year it spawns in relatively sheltered areas, but later in the season prefers more exposed shores. Such a move is accompanied by decreasing wave action and increasing sea temperatures (see also Wourms & Evans, 1974). The increasing temperatures and the resultant greater respiratory demand of the eggs requires that they be subjected to adequate water flow, hence the move to more exposed spawning sites. The spring intertidal spawning habit of this species may serve to isolate it reproductively from Xiphister mucosus (Wingert, 1974), which is only intertidal in the immature stages and spawns subtidally in the winter as an adult. The cottid Clinocottus acuticeps also shows evidence of a preference for a particular type of spawning site. It spawns high in the intertidal zone beneath the alga Fucus distichus. The alga protects the spawn from desiccation and undue rises in temperature when emersed. Spawning does not take place in areas where the alga has been experimentally removed (Marliave, 1981). Another cottid, Enophrys bison, which spawns intertidally is actually a subtidal species. The observation that the egg masses tended to be in certain areas rather than others, suggested that favourable areas, perhaps those most exposed to tidal currents, were being selected over less favourable sites. The advantage that this species obtains from spawning intertidally seems to be threefold. First, embryonic mortality may be lower due to higher ambient oxygen levels during low tides. Secondly, subtidal egg predators can be avoided and thirdly the guardian male could forage subtidally while the eggs were emersed (De Martini, 1978). Another example of intertidal spawning by a subtidal species is the Japanese puffer Fugu niphobles, although in this case the eggs do not remain on the shore, but are washed away by the waves. This species spawns only at sunset on the high waters of rising spring tides (Nozaki et al., 1976; Honma, Ozawa & Chiba, 1980). Further work on the spawning of the grunions (Leuresthes tenuis and L.sardind), well known as species with tidally-synchronized rhythms of egg deposition, has shown that there is a difference between the two. In the Californian grunion, L.tenuis, the spawning runs are governed by the phase of the moon and occur at night whereas in the Gulf grunion, L.sardina, the runs seem to be controlled by tide height and the fish tend to spawn during the day. The latter species’ spawning act is also shorter and is related to the shorter wave periods present in the Gulf (Thomson & Meunch, 1976). In addition, the Gulf grunion’s eggs are smaller than those of the California grunion, a difference which probably evolved in response to the greater regularity of the tidal regime in the Gulf, and the higher probability of hatching within a predicted period (Moffatt & Thomson, 1978). Returning to the resident species, those spawning on rocky substrata probably select a site which requires little preparation apart, perhaps, from claiming a territory around the shelter where the eggs are to to be laid. In most cases this shelter seems to be a hole or crevice which the fish uses as a resting or hiding place. Those species which live on soft substrata, on the other hand, may need to construct a specific nest. In the mudskippers this nest takes the form of a burrow in the mud which the fish may also use as a refuge for the rest of the year. Burrow construction has been described for Boleophthalmus dussumierei (Mutsaddi & Bal, 1969b), Periophthalmus cantonensis (Kobayashi, Dotsu & Takita, 1971) and in greater detail for several other species by Brillet (1969a, 1976). Magnus (1972) has noted that in Periophthalmus kalolo (=P.koelreuteri) the nest

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burrows are restricted to areas below the high water mark of neap tides where they are frequently covered with water and where water leaks into the nest from the surrounding mud when the burrow is exposed. In the goby Pomatoschistus microps, which lives on sand, nest construction consists of selecting a suitably sized bivalve shell, turning it over so that the convex side faces upwards and removing sand from beneath it to form a cavity. The upper surface of the shell is finally covered with sand to prevent dislodgement by waves (Vestergaard, 1976). Having selected or constructed a spawning site the male then has to attract the female to it. In numerous courtship displays the males’ self-advertisement consists of stereotyped movements off the substratum which contrast markedly with the normal movements which are strongly bottom orientated. In the blennies, for example, recent studies have shown the existence of vertical swimming in Blennius rouxi (Heymer & de Ferret, 1976) and vertical looping in Istiblennius zebra (Phillips, 1977b) and in Tripterygion spp. (Wirtz, 1978). Comparable behaviour has been observed in the goby Clevelandia ios (MacDonald, 1975) and in the mudskippers which leap into the air off the mud (Brillet, 1969a,c; 1970). The next phase of courtship consists of behaviour designed to show the female the spawning site, lead her to it and induce her to enter. In blennies vigorous head movements are a characteristic feature of this phase of the display (Abel, 1973; Phillips, 1977b; Heymer & de Ferret, 1976) and they are frequently reinforced by distinctive colour patterns on the head and anterior portion of the body (Zander, 1975). The heads of blennies commonly bear tentacles or crests (Mayer, 1970) which also seem to serve as optical signals, although in Blennius tentacularis at least, they are known to contain sensory cells resembling taste buds and spindleshaped cells which may have a receptor function (Schulte & Holl, 1972). In species which possess tentacles they are present in both sexes but in many cases those of the male become enlarged during the reproductive season (Zander, 1975; Papaconstantinou, 1979). Further optical signals are provided by enlarged dorsal fins in some species (e.g. Blennius sphinx, Zander, 1975) and by distinctive colour patterns (Abel, 1973; Zander, 1975; Losey, 1977; Papaconstantinou, 1979). In addition to the visual stimulation provided by movement, colouration and morphological structures, evidence has accumulated recently to show that some blennies employ olfactory signals to attract others. Initial studies on Hypsoblennius (Losey, 1969) showed that sexually mature non-parental males were attracted by a pheromone produced by other courting or mating males. Females and males guarding eggs were not attracted by the pheromone. This response was considered to promote social facilitation of courtship by attracting males which had not yet mated and possibly to enhance their sexual receptivity. Losey did not investigate the source of the pheromone but suggested that it might originate from glands on the anal fin. Female Hypsoblennius also produce a pheromone which stimulates courtship in conspecific males. This female pheromone probably acts as part of a reproductive isolating mechanism in species whose reproductive behaviour and female colour patterns are very similar (Todd, 1971). Laumen, Pern & Blüm (1974) demonstrated experimentally that the pheromone of male Blennius pavo is produced by glandular tissue on the first two anal fin spines. This pheromone is attractive to females and is produced only by mature males, not by females, immature males or ripe males whose anal fin glands have been removed. The presence of these bulb-like glands on the anal fins of several other species (Blüm, 1972; Zander,

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1972b, 1975; Smith, 1973; de Leo, Catalano & Parinello, 1976; Papaconstantinou, 1979) suggest that pheromone production may be of relatively common occurrence. In male Blennius sphinx the formation of the anal fin gland system can be promoted by injection of mammalian luteinizing hormone (Blüm, 1972). A second type of structure on the anal and/or dorsal fin is also found in blennies. It consists of a fleshy club-like extension of the fin rays (Zander, 1972b, 1975; Smith, 1973; de Leo et al., 1976; Fives, 1980b) and may be an under-developed form of the bulb-like gland (Zander, 1975). Its function is unknown. Once the female has been attracted to the nest site by the male courtship continues. This behaviour is best known in blennies and was summarized in the earlier review (Gibson, 1969a). Further observations on other bleniid species have shown that they all conform to the general pattern previously described (Chasmodes bosquianus: Phillips, 1971a; Alticus saliens and Entomacrodus vermiculatus: Abel, 1973; Blennius rouxi: Heymer & de Ferret, 1976; Hypsoblennius spp.: Losey, 1977; Istiblennius zebra: Phillips, 1977b; Tripterygion spp.: Doak, 1972; Wirtz, 1978; Coryphoblennius galerita: Fives, 1980b). One interesting feature that has come to light is that the courtship display of a male may attract others. This attraction has been observed in Istiblennius zebra (Phillips, 1977b) but the sex of the attracted fish could not be determined. In Trypterygion spp. small males without territories gather around a spawning pair. It was inferred that these “satellite males” may attempt to fertilize the eggs of the spawning female and thereby producing offspring without having to guard them (Wirtz, 1978). The reproductive behaviour of other groups is not as well known as that of the blennies but Martin & Martin (1971) give a brief account of spawning in the clingfish Gobiesox strumosus and MacDonald (1975) mentions courtship behaviour in his account of the biology of the goby Clevelandia ios. Brillet (1969a,c, 1970), however, has recently described the courtship of mudskippers in some detail. It takes place in five phases. The first is a mutual approach, the male erects the dorsal fin, waves the tail and leaps off the substratum. He then circles and butts the female and after a few turns leads her to the burrow. They then enter the nest and when they re-emerge show a variable series of postures and acts reminiscent of the initial sequences of behaviour. Brillet (1970) remarks on the similarity between sexual and aggressive behaviour and that on occasions this ambiguity leads to a transformation from courtship to aggression. Comparable behaviour has also been described in other mudskippers by MacNae (1968, Periophthalmus chrysospilos) and Magnus (1972, P.kalolo). Eventually the female spawns and the eggs are laid on the burrow walls (Kobayashi et al., 1971; Brillet, 1976). This habit of attaching the eggs to the substratum is of almost universal occurrence among resident species and its existence has been confirmed in those which have recently been investigated, e.g. in the Gobiesocidae (Martin & Martin, 1971; Ruck, 1971, 1973b; Shiogaki & Dotsu, 1971, 1972c); in blennioids (Shiogaki &Dotsu, 1972b; Dotsu&Oota, 1973; Ruck, 1973a; Phillips, 1977b; Wirtz, 1978; Fives, 1980b); in the Gobiidae (Brothers, 1975; Shiogaki & Dotsu, 1972a,d); and in the Cottidae (De Martini, 1978). Where the eggs do not adhere to the substratum as in Pholidae and Stichaeidae they are often formed into a ball by the actions of the parents. These elongate forms then curl round the spherical egg mass and prevent it from being washed away. In the case of Xiphister atropurpureus and X.mucosum the guardian parent is always male (Marliave &

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De Martini, 1977). Paternal care of attached egg masses, which are frequently composed of contributions from more than one female, is of very common occurrence (see references to egg attachment above and Keenleyside, 1979) and its significance has been discussed by Ridley (1978) and Blumer (1979). Both authors come to similar conclusions and point out that paternal care is frequently associated with external fertilization and male site attachment. External fertilization ensures that the male has a high probability of genetic relatedness to the offspring he is to guard and site attachment enables the male to attract a succession of females and fertilize their eggs while guarding those from previous spawnings. One possible benefit accruing to females from paternal care is that it allows them to continue to feed, grow, and increase their fecundity. These generalizations are true for the great majority of intertidal residents, but one exception seems to be the mudskipper Periophthalmus cantonensis in which neither parent guards the eggs (Kobayashi et al., 1971). Where maternal care has been observed (e.g. in Centronotus (Pholis) gunnellus, Qasim, 1957) it is associated with poor feeding conditions or viviparity as in the clinid Clinus superciliosus (Veith, 1979, 1980). The latter species may brood as many as 12 batches of eggs simultaneously. It breeds all year round and this, coupled with the territoriality and early maturity of the female, ensures that the greatest number of young are produced by fish located in the most favourable positions (Veith, 1979). The care provided by the male seems to have three basic functions. It protects the eggs from predators, keeps the eggs clean and free from detritus and provides an adequate supply of oxygen for development. Egg predation is prevented by the aggressive behaviour of the male towards intruders (see, e.g., Marliave & De Martini, 1977; De Martini, 1978; Wirtz, 1978); removal of the guardian male results in the eggs being eaten by conspecifics (Wirtz, 1978) or by other predators (De Martini, 1978; Wirtz, 1978). Egg cleaning is performed either by mouthing the eggs, brushing them with the body or fanning. In Hypsoblennius keeping the eggs free from sediment may be the major function of fanning because fanning is governed by water turbulence rather than by oxygen tension (Losey, 1968). In most cases where fanning has been observed it has been assumed that its function is to aerate the eggs but there do not appear to have been any studies on intertidal fishes which have tested this assumption experimentally. Ruck (1973a), however, did note that the eggs of Tripterygion robustum die if the male is removed and adequate aeration is not subsequently provided. Once development has been completed the larvae hatch, a process which has not been well studied, although Denucé (1976) has described the presence of a hatching enzyme in Gobius jozo. In Periophthalmus sobrinus the eggs require a moist atmosphere, but not total immersion in water, to achieve complete development. The larvae hatch very rapidly when submerged by the rising tide (Brillet, 1976). Ruck (1980) also suggest that the hatching of tripterygiid and gobiesocid species which takes place when the eggs are violently agitated or suddenly bathed in cool sea water may represent an adaptation to hatching on a rising tide. This response coupled with a positive phototaxis, may assist the dispersal of the larvae by enabling them to get rapidly into the water column. Larvae which are positively phototaxic on hatching have also been described in Coryphoblennius galerita (Fives, 1980b) and in Periophthalmus sobrinus (Brillet, 1976). In the latter species the photopositive reaction helps the larvae find their way out of the burrows in

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which the eggs are laid. The reproductive cycles of intertidal species follow a general pattern in which there is a fairly restricted breeding season each year. This pattern may, however, be a reflection of the fact that most species which have been studied are temperate rather than tropical forms. Cycles of gonad maturation have been described in the gobies by Healey (1971a, Pomatoschistus minutus; 1972, P.microps), Dunne (1978, Gobius paganellus), Grossman (1979a, Lepidogobius lepidus), in the blennies by Smith (1973, Blennius cristatus), Shackley & King (1977) and Dunne (1977) (Blennius pholis), Lee & Chang (1977, Halmablennius lineatus), Fives (1980a, Coryphoblennius galerita), in the mudskipper Boleophthalmus dussumierei by Mutsaddi & Bal (1970), in the clingfish Gobiesox rhessodon by Wells (1979) and in the stichaeid Xiphister atropurpureus by Wourms & Evans (1974). The environmental factors controlling these cycles are poorly understood although day-length (Healey, 1971a; Wourms & Evans, 1974; Shackley & King, 1977) and temperature (Healey, 1971a; Shackley & King, 1977) are probably of greatest importance. Healey (1971a), however, considered that a simple hypothesis involving only temperature and photoperiod is not sufficient to explain the timing of gonad maturation in Pomatoschistus minutus.

DEVELOPMENT, GROWTH, AND LONGEVITY The life histories of most intertidal fishes are relatively simple and begin with the laying of demersal eggs. After hatching the larvae remain in the plankton for a time which varies according to the species and ambient temperature. Recent accounts of egg and larval development are given for the Gobiesocidae by Ruck (1971, 1973b), Shiogaki & Dotsu (1971, 1972c) and Allen (1979), for the Gobiidae by Shiogaki & Dotsu (1972a,d), for Periophthalmus by Brillet (1976), for the Tripterygiidae by Ruck (1973a, 1980), and for members of the Blenniidae by Shiogaki & Dotsu (1972b), Dotsu & Oota (1973), De Leo et al., (1976) and Fives (1980b). Once development has been completed the larvae metamorphose and settle on the bottom. The factors governing habitat choice at settlement are poorly known, and the little information available is discussed in a later section. Growth patterns after metamorphosis and settlement seem to fall into two basic categories whose characteristics resemble those defined by the r- and K-

selection hypothesis (see Grossman, 1979a; Miller, 1979). The two patterns are typified on the one hand, by early maturity and short life (usually only ≈2 yr) and, on the other, by deferred maturity and longer life during which reproductive effort is spread over several years. Even in the longest-lived species, however, the life span is comparatively short and very few individuals attain an age in excess of 10 yr (e.g., the giant goby Gobius cobitis: Gibson, 1970a; or the shanny Blennius pholis: Dunne, 1977). The divergence in life histories is particularly apparent in the Gobiidae, possibly because most information is available for this family (see review by Darcy, 1980, and Gibson, 1970a; Healey, 1972; Fonds, 1973; Miller & El-Tawil, 1974; Brothers, 1975; Hesthagen, 1975, 1977;

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MacDonald, 1975; Miller, 1975b; Dunne, 1978; Grossman, 1979a; Fouda & Miller, 1981). Much less is known about the growth patterns of fishes from other families (Blenniidae: Dunne, 1977; Fives, 1980a; Gobiesocidae: Wells, 1979; Stichaeidae: Wingert, 1974; Cottidae: Green, 1971a; Tasto, 1975; Chadwick, 1976; Moring, 1979a). The paucity of data on the life histories of all but the commonest species presumably reflects the relative scarcity of intertidal fishes and the consequent difficulties of collecting sufficient numbers to form the basis for a detailed study. As a result of these difficulties, comparative studies of growth patterns are also rare and limited to a few species. In those species which have been examined in this respect, some show little, if any, differences in growth rate from area to area (Gobius cobitis: Gibson, 1970a; Oligocottus maculosus: Chadwick, 1976; Moring, 1979a) whereas others (Blennius pholis: Dunne, 1977; Gobius paganellus: Dunne, 1978) differ considerably. In the last two species the geographic variations in growth rate were attributed to differences in the temperature regime and exposure of the habitats studied. ZONATION AND HABITAT SELECTION While the phenomenon of the differential vertical distribution of intertidal fishes has been recognized for some time, most of the earlier observations were of a basically qualitative nature (see Gibson, 1969a). Some recent studies also record distribution as presence or absence at particular levels (Norris, 1963, for Girella nigricans and Pinchuk, 1976a, for intertidal species in the Kuril Islands) but most workers have lately attempted more quantitative descriptions of distribution. Sasaki & Hattori (1969), for example, studying the ecology of two tide-pool gobies in Japan, caught 90% of the population of one of them, Chasmichthys dolichognathus, in the lower part of the shore whereas the second species, C.gulosus, was distributed more evenly with 60% of the population at the lower levels. Differences in the vertical distribution of four stichaeids and one pholid (Xererpes fucorum) could be distinguished by considering their levels of greatest abundance and maximum vertical height (Horn & Riegle, 1981) and Burgess (1978) found that the green form of X.fucorum generally inhabited higher levels than the red form. Using the median point of the population distribution as a measure of level Critchlow (1972) was able to distinguish between two groups of clingfishes on a platform habitat in the Gulf of California. One group, composed of Tomicodon boehlkei and T.zebra, was found higher than the second group, T.eos and T.myersi. Two other genera of chaenopsid and tripterygid fishes (Coralliozetus and Axoclinus) similarly inhabited different levels on the same shores. Eger (1971) also working in the Gulf of California found two groups of clingfishes. Pherallodiscus funebris and Tomicodon humeralis occurred mostly at high levels, whereas T.boehlkei and Gobiesox pinniger inhabited only the lower intertidal zone. Differential vertical distribution is also found among the eastern Pacific cottids.

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Green (1971a) in an extensive survey of cottid distribution on the coast of Vancouver Island distinguished three categories: those primarily restricted to tide-pools (5 species), those which are found both in tide-pools and subtidally (7 species) and a third primarily subtidal group (4 species). Nakamura (1976a,b) confirmed that Oligocottus maculosus lives mainly in tide-pools and O. snyderi is found in both tide-pools and subtidally. The latter is also a low level species in California where larger individuals and two other cottids, Clinocottus analis and Artedius lateralis, occupy similar levels. The main difference in the distribution of the three species is that smaller individuals of Clinocottus analis are found higher on the shore than both larger ones and all sizes of Oligocottus snyderi and Artedius lateralis (Yoshiyama, 1980). Some evidence for differential size distribution within a species’ vertical range was also found on the Atlantic coast of France where three groups of fishes could be distinguished; one found over the whole intertidal zone, another restricted to the upper levels and a third found only at the lowest intertidal levels and sublittorally (Gibson, 1972). Zander (1967) similarly describes three groups of salariine blennies living at high, intermediate, and low levels on rocky shores in the Red Sea, the highest living species, Lophalticus kirkii being found mainly out of water in the supralittoral zone. Two gobies, Acentrogobius ornatus and A.meteori, living in the same region were limited to the mid-littoral and upper sublittoral zones on sand or in sand-filled gullies. In the virtually tideless Mediterranean where fishes are not subjected to long periods of emersion, several studies on the Blenniidae have demonstrated that here, too, differences in the depth distribution of individual species can be recognized. The vertical distribution is mainly dependent upon tolerance to wave action, although light intensity (Zander & Heymer, 1976) is also important. Three groups can be distinguished; those inhabiting the “surf zone”, those of the “turbulence zone”, and those preferring deeper, less turbulent water (Zander, 1972a). Such zonation may be quite precise even within a depth range of about 1 m as Zander (1980) found in the Adriatic. Here Lipophrys (Blennius) adriaticus was found mainly in the first 25 cm below the surface, L.canevai from 25–100 cm and L. dalmatinus from about 50 cm downwards, although within these depth ranges the actual microhabitats occupied were different. Comparison of the depth distribution of species in areas of different exposure to wave action suggested that in the great majority of cases the minimum depth at which they were found increased with exposure (Zander, 1972a). Critchlow (1972) also found differences in the vertical distribution and abundance of clingfishes according to exposure to the Gulf of California and on the French coast the abundances and upper limits of certain species were noted to be markedly lower on exposed when compared with sheltered areas (Gibson, 1972). Apparently then, in many cases intertidal fish on rocky shores respond to wave action by shifting their distribution downshore, in contrast to sessile invertebrates and algae which extend their range upwards on exposed coasts (Newell, 1979). One exception has, however, been described. Oligocottus maculosus has a higher lower limit on exposed when compared with sheltered shores, a feature which is considered to be related to a need for a minimum period of low turbulence for survival (Green, 1971a). The depth distribution of fish migrating into the intertidal zone or sandy or muddy shores has not been as extensively studied as that of the residents on rocky shores. The apparent uniformity and lack of physical niches on such shores might be expected to

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favour absence of zonation among tidal migrants, but a study of one beach demonstrated that species were zoned according to depth and that within a few species there was a correlation between depth and fish length (Gibson, 1973a). Where species overlap in their vertical distribution there is considerable evidence that the actual microhabitats occupied by each are different. Eger (1971), for example, found that the clingfishes he studied in the Gulf of California differed in their preferred substratum; some were found most abundantly on cobble beaches, others on boulder or pebble beaches. Critchlow (1972) describes an extreme example of habitat partitioning where the clingfishes Tomicodon eos and T.myersi are obligate associates of seaurchins which live in rock pockets. T.eos is found on the lateral walls of the pockets but T.myersi only on the bottom of the pockets beneath the seaurchin itself. Two other species, T.boehlkei and T.zebra, are also found on the lateral pocket walls at low tide but leave during high tide and it was calculated that they only overlap spatially with the obligate associates 64% of the time. Critchlow’s (1972) general conclusion was that in habitats where species diversity is high there is greater microhabitat specialization. Other examples of habitat partitioning are given by Gibson (1972) for fishes on the French coast. There the “typical habitats” of several species were examined with respect to the level, surface area and degree of cover in the pools in which they were found. All species were found to be different in their habitat requirements. Marsh, Crowe & Siegfried (1978) determined that some measure of cover was the best predictor of the species richness, abundance, and biomass of South African tide-pool clinids and Burgess (1978) showed that at low tide the red and green colour morphs of Xererpes fucorum almost always occupy algal substrata which match their body colouration. The three colour phases of the penpoint gunnel, Apodichthys flavidus, similarly show a general tendency to inhabit algae equivalent in colour to that of their body (Wilkie, 1966). Clinocottus analis also prefers plant cover in the form of the red alga (Pterocladid) and the surf grass (Phyllospadix) even though these are not dominant in its habitat (Mollick, 1968) and the distribution of another sculpin, Oligocottus snyderi, is well correlated with the abundance of macrophytes (Green, 1971a). The gobies inhabiting the brackish lagoons on the coast of Corsica (Casabianca & Kiener, 1969) and in bays in California (Brothers, 1975) can likewise be separated according to their degree of preference for vegetation. Wheeler (1980) has recently reviewed the relationships between fish and algae in temperate waters. Finally, it is worth mentioning that, for practical reasons, the majority of the data on vertical distribution and microhabitat occupation result from observations made at low tide. Relatively little has been discovered about the changes in distribution that may occur on the flooding tide. Although many species are known to move about the intertidal zone when it is submerged the extent and direction of these movements are poorly documented. This aspect of behaviour is discussed in a later section. The instances of differential vertical distribution given above are sufficient to demonstrate that zonation of intertidal fishes is a real phenomenon. The mechanisms whereby such zonation patterns are established and maintained are poorly understood and recent studies have only just begun to examine some of the problems. Establishment of zonation begins with the larvae, although in a wider sense it may be considered to begin with the egg because the great majority of resident intertidal species

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lay attached demersal eggs in the adult habitat. Such behaviour reduces the degree of dispersal of premetamorphic stages as well as the chance of predation in the egg phase. In a few cases, dispersal of the young is greatly minimized by the adoption of viviparity, as in some clinids for example, so that the young are released in the direct vicinity of the adults. Larvae are planktonic, however, and the mechanism whereby they are recruited back into the intertidal zone remains for the most part unknown. Norris (1963) has suggested that thermotaxis and association with floating debris may be involved in the onshore migration of prejuveniles of Girella nigricans and Ehrlich, Hood, Muszynski & McGowen (1979) point out that selection of warm water by larvae of Hypsoblennius gilberti will help them to remain near-shore and aid settlement intertidally. Creutzberg, Eltink & van Noort, 1977) have postulated that plaice (Pleuronectes platessa) larvae use the flood tidal currents to enter the extensive intertidal zone in the Dutch Wadden Sea. Starvation releases pelagic swimming, whereas feeding inhibits this tendency. The tidal flats where food is much more abundant than offshore thus act as a ‘trap’ for settling larvae. A similar use of tidal currents for immigration into the nursery grounds has been suggested for larvae of the stone flounder (Kareius bicoloratus) in Japan (Tsuruta, 1978). Whatever mechanism the larvae employ to reach the intertidal zone, once there, zonation could be the result of one or a combination of settlement strategies. Settlement could be randomly or uniformly distributed over the whole zone and zonation established by differential mortality; the greatest mortality occurring outside the favourable zones. Alternatively, settlement could be localized as a result of active site selection by the larvae which, once a preferred site has been selected, transform into the benthic juvenile stage. Unfortunately, information which would enable these or alternative hypotheses to be tested is almost entirely lacking. Some indication that settlement may not be random in Oligocottus snyderi and O.maculosus was obtained when it was observed that postlarvae of the former species, which inhabits low intertidal levels as an adult, were only found in low level pools. Postlarvae of O.maculosus, on the other hand, were found mainly in upper tide-pools, the habitat of the older stages (Nakamura, 1976b). Stephens, Johnson, Key & McCosker (1970) also found that the settled larvae of Hypsoblennius were only present in the habitat where the adults normally spawned. The only study of the substratum preferences of settling larvae is that carried out by Marliave (1977) who presented the larvae of six species with a variety of substrata on which to settle. In five species out of the six one substratum was preferred over all others. The species which showed no preference for any one substratum type (Artedius lateralis) also inhabits a wide diversity of substrata as an adult. In the five species which did show preferences, the substrata chosen for settlement corresponded, although not always exactly, to those of the adult. Differences between the substrata selected by the larvae and those in the habitats occupied by the adults could be explained by factors relating to body size because Marliave (1977) found that structural characteristics rather than chemical composition were responsible for the results of his experiments. Interstitial spaces accommodating larvae are no longer adequate for juveniles, for example, and the rough texture of rocks, although allowing adhesion by adult clingfish, are not suitable for the younger stages whose small suction discs are only capable of adhering to smooth surfaces. In addition to substratum selection, other factors such as salinity and water movement were considered to be important in the final determination of habitat choice.

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Although a change in microhabitat may take place as the fish grow, the later stages of resident species still have to ensure that they maintain their position in the finally selected zone rather than wander at random over the intertidal zone or leave it altogether. The few studies which have examined habitat maintenance in adults all deal with rocky shore species. Wilkie (1966) was able to demonstrate that Apodichthys flavidus, which varies in colour from red to green, can select vegetation of a colour matching that of its body when given a choice. Such a selection is made only when no other cover is available, however, and when rocks are present it prefers to hide beneath these, irrespective of the colour of algae attached to them. It was concluded that plant cover is probably only utilized at specific times, such as while feeding, and the utilization of a substratum which matches the body colouration may act as a defence against predation in the relative insecurity of algal patches. When threatened or resting the fish hides under rocks. Another pholid with two colour phases, Xererpes fucorum, only selects matching algae when emersed (Burgess, 1978). Under water this species does not show such a preference, although it will select algal cover (regardless of colour) in preference to bare rocks and bare rocks are preferred to areas with no cover. The interpretation of these experimental results was that X. fucorum prefers different substrata according to tidal conditions. It occupies matching algae at low tide but moves to under rock habitats at high tide, perhaps as a protection against predation and turbulence. This interpretation was supported by observations made while skin-diving at high tide when no fish were seen among algae. The presence of cover is also an important factor governing the choice of habitat by Clinocottus analis. This species selects densely matted vegetation or the dark spaces between rocks rather than open or flat areas of any shade (Mollick, 1968) and in artificial pools tends to occupy those with the greatest amounts of cover (Rickhus, 1981). The selection of a particular substratum is made entirely visually rather than by the use of chemical or tactile cues and the presence of food organisms plays no part (Mollick, 1968). The blenny, Chasmodes bosquianus, also shows a preference for cover rather than open spaces (Phillips, 1977a). Nakamura (1976a) found experimentally that preference for vegetation strongly influences the distribution of Oligocottus snyderi, but superimposed upon this preference is the need to avoid adverse temperatures high on the shore. Thus, although habitats suitable in terms of plant cover are available at upper levels they are not inhabited by O. snyderi because it is unable to cope with high temperature extremes and actively avoids them. Its congener, O. maculosus, on the other hand, selects rock substrata in preference to eel grass or open areas when presented with a choice. Furthermore, in artificial tide-pools it will select the shallow end of the depth range available whereas O. snyderi shows no such depth selection. The difference in the distribution of the two species can, therefore, be explained to a large extent by their respective choices of substratum and depth and their behaviour in relation to temperature. The absence of social interaction between the two species meant that the results of the selection experiments were the same whether the two species were tested together or separately. Stephens et al. (1970), however, working with Hypsoblennius jenkinsi and H. gilberti found that habitat preferences shown by the species in isolation were strongly restricted and much more distinct in the presence of a conspecific. The finding common to each of these studies is that cover of some description must be present if one habitat is to be selected in preference to others. This necessity for the

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presence of cover, together with their strong thigmotactic tendencies, may be used to explain the limitation of rocky shore species to their particular habitat and why they are rarely, if ever, found elsewhere. Other than this no general statement can be made regarding the mechanisms involved in habitat selection by intertidal fishes. The work by Nakamura (1976a) and Marliave (1977) quoted above is certainly a step in the right direction but much more requires to be done along these lines before the state of knowledge attained for other particular groups, e.g., gastropods (Underwood, 1979) or even the invertebrates in general (Newell, 1979) can be approached.

INTERTIDAL MOVEMENTS EXTENT OF MOVEMENT In order to satisfy their requirements for an adequate food supply many species have to make extensive foraging excursions away from their low tide locations. In the great majority of cases these feeding excursions are made at high tide when conditions are most favourable. Although feeding may be possible when emersed, only those forms which have evolved a semi-amphibious mode of life can do so efficiently. For the remainder movement at high tide is necessitated by their limited powers of terrestrial locomotion as well as the need to avoid desiccation hazards and the risk of predation by those birds or terrestrial animals which can only enter the intertidal zone when the tide is out. The food organisms themselves, of course, may also be more readily available at high tide. The extent of these excursions may be related to the abundance of food; Newell (1979) has suggested for example, that invertebrates on sandy shores make much more extensive movements than those on rocky shores and that where food is abundant, as at lower levels on rocky shores, it is partitioned among the inhabitants by rigid territorial mechanisms. Studies of fish movements on sandy shores have indeed shown that they are considerably more extensive than on rocky shores, but the difference is probably related more to the fishes’ inability to survive intertidally than to differences in food abundance. The extent of movement of fishes on sandy beaches should, therefore, be compared with the tidal visitors to rocky shores rather than with the rocky shore residents. The movements of fishes on open beaches are relatively easy to discern because sampling at different states of the tide is simpler than it is in rocky areas. MacNae (1968), Kobayashi, Dotsu & Takita (1971), and Brillet (1975) have described the movements of several species of mudskippers which follow the tideline as the water advances and recedes, but most other species do not possess the amphibious capabilities of the mudskipper and remain submerged. Several studies have demonstrated that on sandy shores in particular, the movement of fishes into the intertidal zone on a rising tide is extensive and well synchronized. The movements of young flatfish, for example, are particularly well documented as a result of observations using underwater television (Tyler, 1971), netting (Kendall, 1966; Gibson, 1973a; Kuipers, 1973; Lockwood, 1974) or direct observation (Gibson, 1973a, 1975, 1980). Such tidal visitors avoid being stranded by leaving the shore on the ebb tide and in contrast to residents do not have to contend with the problems of selecting a low tide refuge where they can await the next

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period of immersion. Return or fidelity to particular areas on a beach might not be expected, therefore, but Riley (1973) in a series of tagging experiments with young plaice (Pleuronectes platessa) found that they are restricted in their movements (<0.5 km over several months) and that they will return to a particular location if displaced. Movements on sandy shores are primarily feeding migrations which enable fishes to utilize food resources not available to them at low tide, although such movements may also serve the function of predator avoidance. Similar movements are known for fish populations of tidal creeks (e.g., Cain & Dean, 1976) and for tidal visitors to rocky shores (e.g., Gibson, 1972; Thomson & Lehner, 1976). Inherent in the movement of a resident species away from a specific zone or low tide location, however, is the danger that it might not regain its position by the next low tide unless some form of behavioural restriction on movement is imposed. Consequently, all studies of the high tide movements of intertidal residents have shown that they are limited in extent compared with those theoretically available for movement. In addition, many species have evolved homing mechanisms which enable them to return to favourable low tide locations. Examples of very restricted movement on rocky shores are given by Abel (1973) for Alticus saliens in Ceylon which rarely moves more than 1 m from its resting place and Critchlow (1972) for the clingfish Tomicodon eos and T. myersi which are restricted to sea-urchin pockets. Other workers simply note that movement does take place without detailed descriptions of its extent (Gibson, 1972; Chang, Lee & Wu, 1977; Phillips, 1977b; Burgess, 1978). Green (1971b) in a detailed study of the movements of Oligocottus maculosus found that the percentage of fish leaving pools at high water depended on both temporal and spatial factors. In exposed areas this species does not regularly leave the pools at high water. In sheltered areas fish do leave pools, but only the young fish show a complete shift in distribution over the high tide period. The proportion of fish leaving pools also varies with the time of year, the proportion being greatest in the summer. The situation is further complicated by the fact that there are considerable differences in the degree of movement shown by individual fish. Richkus (1978) examined the movements of another cottid, Clinocottus analis in an area of 1000 m2 on the Californian coast over 5 months and also found that many fish moved from pool to pool. Most, however, were recaptured in pools close to the original tagging point. In this respect Richkus’ (1978) findings were very similar to those described earlier for the European Blennius pholis (Gibson, 1967b). Most movement studies, however, have concentrated on an examination of the fidelity of fish to a particular pool or general area with the implication that if fidelity is high, movement must be limited in extent. Richkus’ (1978) study of Clinocottus analis demonstrated that some individuals can be found in the same pools on successive occasions for as long as 20 weeks but the general conclusion was that this species shows fidelity to a general and limited area rather than particular pools. Several factors were found to affect pool residence times, particularly fish length and pool stability. Larger fish had a higher probability of recapture than smaller ones, suggesting that colonization of new areas is performed by juveniles and that fish become more sedentary with age. Such a finding could, however, be explained by differential homing abilities as demonstrated for Oligocottus maculosus (Craik, 1978, 1981, see p. 394). High fidelity to specific pools is also shown by the latter species

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(Green, 1971b) and Clinocottus globiceps (Green, 1973) in British Columbia because 86% and 94% of these two species, respectively, were found in their initial capture pool after more than 2 weeks. Moring (1976), however, only recaptured about 7% of marked O. maculosus on a shore in northern California and Craik (1978) showed that the degree of fidelity shown by this species depends upon the amount of turbulence present and the regularity of the terrain. Fidelity is greatest in regions of high turbulence and irregularity. In groups other than cottids, Marsh et al., (1978) demonstrated that South African clinids moved between pools at high tide and that some individuals showed fidelity to particular pools over a 4-month period. Burgess (1978), on the other hand, found that only about half of the pholid Xererpes fucorum were recaptured after one tide. Sasaki & Hattori (1969) provide data which indicate a difference in residence times of two species of the gobiid genus Chasmichthys. In C. dolichognathus, a low level species, an average of only 16% of the original inhabitants were recaptured in the same pool after 2 months. The higher level C. gulosus, in contrast, had a 60% recapture rate over the same period. In both species, individuals living higher on the shore tended to stay longer in pools than those at lower levels, possibly because the latter were immersed for a greater proportion of the time and consequently had a higher probability of dispersal. This fidelity to a particular area and the restricted movement within it is reflected in the relatively slow recolonization rates of pools which have been experimentally depopulated. Bussing (1972), for example, poisoned a supralittoral pool on the Marshall Islands and re-sampled it after 3 weeks. After this time interval the species diversity was lower and the number and biomass had only reached about 50% and 3%, respectively, of their original levels. Initial recolonization was due primarily to young individuals which explains the much lower second biomass value. Marsh et al. (1978) similarly found that the mean length, number and weight of clinids were lower 12–24 weeks after removal of the original inhabitants of the pools and that recolonization was performed mainly by younger fish. Thomson & Lehner (1976), however, found that changes in the tide-pool populations they studied in the Gulf of California were affected more by short term seasonal temperature effects than by the effect of successive recolonization following defaunation at intervals exceeding 4 months. HOMING AND ORIENTATION Allied to the observations that many fishes move over restricted areas and show fidelity to specific pools or particular areas is the implication that they are able to recognize their surroundings, remember them and use this topographic memory to return to their original low tide location. This process, generally termed homing, has been investigated in several species. Before homing (defined by Gerking (1959) as “the return to a place formerly occupied instead of going to other equally probable places”) can be said to take place, however, it must be shown that movements towards “home” are directed. Consequently, although the presence of an individual in a particular pool on successive occasions is suggestive of homing, such an observation cannot be regarded as proof. The best evidence for the existence of homing behaviour comes from experiments in which fish are displaced and subsequently recaptured in their original location. Since the earlier studies by Williams (1957) and Gibson (1967b) such displacement experiments have

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mainly involved Oligocottus maculosus, but homing has also been recorded in Hypsoblennius gilberti (Stephens et al., 1970), Clinocottus globiceps (Green, 1973), and Tripterygion (Heymer, 1977) although not in stichaeoids (Barton, 1973). Both Green (1973) and Heymer (1977) found that ≈60% of transplanted fish returned to their original point of capture. In the case of Clinocottus globiceps the figure was much higher (88%) if the percentage of fish actually recovered is considered. A very similar homing percentage was obtained with Oligocottus maculosus when displaced over distances of up to about 100 m. Subsequent studies confirmed the homing abilities of O. maculosus (Khoo, 1974; Craik, 1978, 1981) and showed that homing success is age, and hence length, dependent. Younger, smaller fish are much less able to home than older fish when displaced, but some loss of homing ability is evident in the oldest individuals (Craik, 1981). All ages, however, show a similar ability to locate a particular pool within their home range. The results of Green’s (1971b) studies led him to conclude that homing may not depend upon prior knowledge of the area traversed while homing, because (1) homing occurs at times of the year when the fish are seldom, if ever, out of their home pool and, (2) that out of pool movements, when they do take place, are not as great as the experimental transplant distances. If spatial familiarity is used in homing then it must have been gained at an earlier age when fish show no great fidelity to particular pools. Craik (1978, 1981) examined this latter suggestion and found evidence to suggest that the youngest fish do move over wide areas and have the opportunity to acquire spatial information, but later adopt a home pool. It was also shown that after periods in captivity these young fish are unable to retain their homing ability as long as older fish which can still home after 6 months away from their natural habitat. The environmental cues employed by O. maculosus while homing remain unknown, but vision and olfaction are two processes which have been shown to be important if an individual is to retain its homing ability. Blinded fish and those with their olfactory organs destroyed are less able to home than normal fish over distances up to 400 ft (121 m) and olfaction seems to be the more important of the two senses (Khoo, 1974). Craik (1978) confirmed the rôle of vision and olfaction in homing but found, in addition, that whereas one sense is sufficient to allow homing in older fish both are necessary in the younger ones. She also demonstrated that neither conspicuous landmarks nor olfactory cues emanating from the home pool are recognized. Summarizing the results for O. maculosus, the evidence for homing (using Gerking’s, 1959, definition) is strong, and familiarity with an area is gained while young and remembered for the rest of life. Vision and olfaction play an important part but the problem of the actual cues used in navigation back to the home pool remains essentially unsolved. Evidence that another species, Bathygobius soporator, uses topographical cues in its orientation was obtained by Aronson (1971). Earlier studies (Aronson, 1951, 1956) had shown that this species acquires its knowledge of an area by swimming over it at high tide and that at low tide it can jump accurately from one pool to another. Further experiments in artificial tide-pools (Aronson, 1971) confirmed the earlier hypothesis and showed that the fish can retain the memory of an area for at least 40 days. In addition, it was demonstrated that fish from sandy shores, which had no prior experience of rocky areas, also improved their jumping accuracy after swimming over the test area. This improvement was not as marked as in those individuals naturally inhabiting rock-pools,

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however, and the difference between the two types suggests a genetic influence on leaping ability. Landmarks (piles of shells) placed in the artificial pools in different positions provided some insight into the kind of cues that the fish might be learning at high tide in nature because it was found that they can learn to orient to notches in the pool walls but not to mounds. Decisive experiments as to the function of homing have yet to be performed but at least two suggestions have been put forward. Williams (1957) proposed that homing in intertidal fishes enables them to avoid being left by the tide in unfavourable situations. Green (1971b) questioned the advantages of such a function and postulated that, for Oligocottus maculosus at least, homing may serve the same purpose as territoriality. That is, it acts as a dispersal mechanism thereby preventing collection of large numbers of individuals in the most favourable areas and hence decreasing the probability of localized destruction of populations by adverse environmental conditions. TERRITORIALITY AND AGGRESSION Territoriality is a further behavioural mechanism which is linked to the phenomenon of restricted movement in the sense that it maintains individuals in a limited area for considerable periods. Non-reproductive territoriality has been described in many species of intertidal fishes from most of the main families but there is considerable variation between species in the degree to which territorial behaviour is exhibited. Critchlow (1972), for example, found among the clingfishes that Tomicodon myersi preferred, or at least tolerated, close contact with conspecifics, T.boehlkei and T.zebra maintained individual distances, but that T.eos was highly territorial in both sexes. Brothers (1975) also discovered differences between three gobies inhabiting intertidal mudflats. Clevelandia ios is non-territorial Quietula y-cauda is territorial but only in males, whereas all adults of Ilypnus gilberti are highly aggressive and territorial. Another symbiotic burrow-dwelling goby (Lepidogobius lepidus: Grossman, 1979b) does not possess fixed territories, but rather defends a “personal space” around itself (Grossman, 1980). Miles (1974) also demonstrated the presence of agonistic behaviour in Californian gobies and in a laboratory study of Gobiosoma chiquita found that the males maintain territories and set up a dominance hierarchy based on size. Size, together with past experience and prior residency are also factors which affect the outcome of agonistic encounters in Lepidogobius lepidus (Grossman, 1980). Mudskippers, too, are known to be territorial (Brillet, 1969b, 1975; Magnus, 1972) and Periophthalmus sobrinus in Madagascar may occupy either of two types of territory. One is centred on a single burrow and the other consists of two parts either linked by a subterranean passage or separated into a low tide (feeding) or high tide (resting) territory (Brillet, 1975). Territoriality and agonism also seem to be common among the blennies and are known from earlier studies (Gibson, 1968; Losey, 1968; Stephens et al., 1970). Recent field work on the Hawaiian rockskipper Istiblennius zebra has shown that only males maintain territories although females do show agonistic tendencies. This species does not seem to be extremely competitive as regards space and vacated territories remain unclaimed for relatively long periods (Phillips, 1977b). Another field study by Nursall (1977) on the Caribbean Ophioblennius atlanticus revealed that both sexes are permanently territorial

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and that intruders planted in an area already occupied are rarely successful in taking over residents’ territories. Conversely, if some residents are removed, the neighbours rapidly take over the vacated space. Aggressive behaviour and dominance hierarchies are also found in the behavioural repertoire of the striped blenny Chasmodes bosquianus (Phillips, 1971a,b, 1974; Phillips & Swears, 1979) and the crested blenny Hypleurochilus geminatus (Lindquist & Chandler, 1978). In the latter species evidence was obtained that the smaller individuals could protect themselves from attack by selecting small refuges. Territories are defended in aquaria by Lophalticus kirkii but not by Istiblennius edentulus (Zander, 1967). No evidence for territorial defence by Lophalticus kirkii was found in nature, however, (Zander, 1967), and this observation highlights the problem of extrapolating laboratory studies to the field. In some studies which combine both types of observation, laboratory results have generally been confirmed but the frequency of agonistic behaviour is usually lower in the field (e.g., Miles, 1974, for gobies; Ordzie, 1969, for Girella nigricans). An explanation for such findings may be that social structure in nature is in a state of dynamic equilibrium and agonistic behaviour and territorial disputes are at a minimal level sufficient to maintain the status quo. Such an explanation receives some support from Miles’ (1974) observation that the rate of aggressive activity was increased when the existing social structure was altered by introducing new fish. Finally, even though widely studied, the only group of intertidal fishes so far known to show no signs of agonistic of non-reproductive territorial behaviour are the cottids. They can often be found in groups resting against or even on top of one another (Mollick, 1968; Green, 1971b; Nakamura, 1976a; Richkus, 1978). Such aggregations, however, may be the result of a positive attraction between individuals (Richkus, 1981) rather than the absence of aggression. An interesting physiological consequence of aggressive tendencies has been described in Blennius pholis. This species which is known to exhibit agonistic behaviour, at least in aquaria (Gibson, 1968), shows a decreased growth rate when kept in visual and olfactory contact with a conspecific (Wirtz, 1974, 1975). The oxygen consumption also increases when the fish confronts its mirror image due to an increase in metabolic rate and activity (Wirtz & Davenport, 1976). Wirtz & Davenport hypothesized that these effects are a manifestation of a deviation from a preferred social environment and that B.pholis is essentially a solitary species. The functions normally ascribed to non-reproductive territoriality are partitioning of the environment to ensure efficient use of resources such as food and shelter and consequently as a means of controlling population size. Several studies have shown that territory holders or dominant individuals have greater access to shelter and favourable areas than subordinates and this was the general conclusion reached by Brillet (1975) after his studies on Periophthalmus sobrinus. In this species, and in P.koelreuteri (Brillet, 1969b), the population consists of two “clans”, dominant fish with territories and subordinates without, the subordinates being forced to occupy the less favourable areas of the shore. Priority of access to food and shelter, particularly when limiting, has also been described for dominant individuals of Lepidogobius lepidus (Grossman, 1980), Gobiosoma chiquita (Miles, 1974), and Chasmodes bosquianus (Phillips, 1971b; Phillips & Swears, 1979). Access to shelter is especially important for small intertidal fishes which are incapable of sustained swimming. They can best avoid capture by predators by

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entering holes and crevices which are too small for the predator to follow. In one species, Chasmodes bosquianus, it has been demonstrated experimentally that dominants escape predatory attack more frequently than subordinates (Phillips & Swears, 1979). Removal by predators of individuals without access to shelter thus acts as a means of population control. A similar suggestion was put forward by Marsh et al. (1978) to explain their observation that aggression occurs between juveniles and adults of Clinus superciliosus and that the smaller fish are forced to occupy open areas where they are probably more vulnerable to predation. Subordinate individuals of Ophioblennius atlanticus, on the other hand, are able to take up small “interstitial” territories between those of full territory holders and these interstitial territory holders seem to provide a reserve for the maintenance of optimal population levels (Nursall, 1977). RHYTHMIC BEHAVIOUR For intertidal animals which live in an environment where regular and predictable changes in conditions are the rule, synchronization of activity with such changes is clearly of importance. Synchronization may be brought about by a direct response to changing conditions such as the flooding of a pool and subsequent alterations in temperature, salinity or light intensity but numerous recent studies have shown that many species also possess endogenous tide-related rhythms of behaviour. The endogenous nature of this rhythmic behaviour has been demonstrated by recording locomotory activity under constant conditions in the laboratory where the fish show alternating bouts of rest and activity with the periods of activity occurring at or near the time of high tide. Such tidal rhythms seem to be of common occurrence because they have been recorded in representatives of many different families (Table II) and more examples may be expected to come to

TABLE II Known examples of endogenous tidal rhythmicity in inter tidal fishes: see p. 364 for definition of ecological category

Family and species

Habitat

Ecological category Reference

Blennius pholis

Rocky shores

Resident

Gibson, 1967c, 1971

Coryphoblennius galerita

Rocky shores

Resident

Gibson, 1970b

Blennius cristatus

Rocky shores

Resident

Stahl, 1973

Rocky shores

Resident?

Doak, 1972

Rocky shores

Resident

Green, 1971c

Blenniidae

Tripterygiidae Tripterygion capito Cottidae Oligocottus maculosus

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Gobiesocidae Tomicodon humeralis

Rocky shores

Resident

Gollub, 1974

Gobiesox pinniger

Rocky shores

Resident

Gollub, 1974

Gobbiidae Pomatoschistus minutus

Sandy beaches Tidal visitor

Gibson & Hesthagen, 1981

Muddy shores Resident

Ishibashi, 1973

Periophthalmidae Boleophthalmus chinensis

Periophthalmus cantonensis Muddy shores Resident

Nishikawa & Ishibashi, 1975

Pleuronectidae Pleuronectes platessa

Sandy beaches Tidal visitor

Gibson, 1973b, 1976

Platichthys flesus

Sandy beaches Tidal visitor

Gibson, 1976

Peltorhamphus latus

Sandy beaches Tidal visitor

Roper, 1979

Rhombosolea tapirina

Sandy beaches Tidal visitor

Roper, 1979

Sandy beaches Tidal visitor

Gibson, 1976

Estuaries

O’Connor, 1972

Bothidae Scophthalmus rhombus Soleidae Trinectes maculatus

Tidal visitor?

light in the future. Descriptions of the rhythms of those species known to exhibit such behaviour have been given elsewhere (Gibson, 1978; see also Schwassman, 1980) and need not be detailed here. Endogenous rhythms operating independently of the environment are usually considered to enable an animal to be prepared for predictable change. Such preparation may be particularly important for intertidal species whose opportunities for feeding or mating may be limited to a few hours each tidal cycle or even each 24 h if they are strongly diurnal. On occasions, however, environmental conditions may be unsuitable for normal activity and the underlying inherent rhythm may be overridden by a direct response to prevailing conditions. Green (1971c), for example, has shown that Oligocottus maculosus possesses an endogenous tidal rhythm of locomotion in which maximum activity is phased with high tide, but that at times of high turbulence in the field it may be totally inactive at high water. In cases such as this the endogenous rhythm may function as a clock which keeps in phase with the tidal cycle so that when favourable conditions return the fish is capable of immediately resuming its normal activities at the correct time. Specific behaviour patterns or local tidal conditions may also have an effect on the phase relationships of rhythmic activity. In those rocky shore species which show rhythmicity the activity peaks are usually centred on the predicted time of high tide whereas in species from sandy shores the peaks occur on the ebb tide. Although the adaptive significance of such a phase difference has not been investigated experimentally

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it may be related to the different movement patterns of the two groups. The ebb-related activity of young flatfish, for example, is seen as a mechanism which ensures that downshore movement is correctly phased and prevents stranding of species which are unable to survive intertidally. Species on rocky shores, however, do remain intertidally at low tide and an ‘anti-stranding’ device is not necessary. For these species it may be more important to synchronize their activity with the high water phase of the tidal cycle when they can move and feed freely. In areas where the tidal range is small and fluctuations in water level are often more dependent upon meteorological conditions fish seem to regulate their activity according to the light: dark cycle rather than to the tides. In the Baltic and its approaches the gobies Gobius niger (Hesthagen, 1976) and Pomatoschistus pictus (Hesthagen, 1980) show no evidence of tide-related behaviour and laboratory studies of two Mediterranean blennies have suggested that they are active at dawn and dusk (Gibson, 1969b). Studies of representatives of species which live in both tidal and non-tidal areas e.g., Pomatoschistus minutus (Gibson & Hesthagen, 1981) and Platichthys flesus, (Muus, 1967; Gibson, 1976) have also shown that their rhythmic behaviour reflects the relative importance of the tidal regime in their particular environment. Even in areas where tidal fluctuations may be considerable, however, several field studies have demonstrated that many species are strongly diurnal, drastically reducing or ceasing activity at night. Such diurnal changes in behaviour have been recorded in mudskippers (Magnus, 1972; Gordon, Ng & Yip, 1978), clingfishes (Critchlow, 1972), the clinid Mnierpes (Graham, 1973), the pholid Apodichthys flavidus (Wilkie, 1966) and the blennies Ophioblennius atlanticus (Nursall, 1981) and Blennius sanguinolentus (Taborsky & Limberger, 1980). The activity pattern of the last species in the field (Taborsky & Limberger, 1980) differs markedly from that recorded in natural light cycles in the laboratory (Gibson, 1969b). The reason for the discrepancy is unresolved but it may be that the laboratory pattern is an artefact of captivity or that there is a real difference between the behaviour of the fish studied by Taborsky & Limberger (1980) in the Adriatic and those in the western Mediterranean (Gibson, 1969b). Other laboratory studies on the blennies Blennius pholis (Gibson, 1971) and Coryphoblennius galerita (Gibson, 1970b) suggested that they are also diurnal because darkness greatly depresses their activity. The presence of diurnalism in the great majority of intertidal residents is probably related to their dependence on sight for carrying out most of their activities, particularly the detection of predators and for feeding.

FOOD AND FEEDING The change in emphasis from a descriptive to an analytical approach referred to on p. 363 is most apparent in recent studies of the feeding ecology. Although basic descriptions of the diet of individual species form the core of such studies many workers have attempted to extend their findings by making inter- and intraspecific comparisons of diet in relation to competition and distribution. Straightforward descriptions of diet or general feeding habits are given for

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mudskippers by Berry (1972) and Sarker, Al-Daham & Bhatti (1980), for gobies by Casabianca & Kiener (1969), Healey (1972), MacDonald (1975), Miller (1975a) and Darcy (1980), for blennies by Smith (1973), Lee & Chang (1977) and Lindquist & Chandler (1978), for Clinus cottoides by Christensen (1978a) and for Gobiesox rhessodon by Wells (1979). Basic dietary surveys of communities have been carried out by Chang & Lee (1969) and Chang, Lee, Lee & Chen (1973) in Taiwan, Zander & Heymer (1976) in the Mediterranean, and Thomson & Lehner (1976) in the Gulf of California. The last authors divide the intertidal fish community into grazers, planktivores, benthic invertebrate predators, and piscivores. Numerous other authors have taken a similar approach in partitioning communities into feeding types either on the basis of their feeding behaviour (Critchlow, 1972), the type of food eaten (Porumb, 1968; Gibson, 1972; Barton, 1973; Chang et al., 1973), brain and sense organ structure (Bath, 1965), gut morphology (Mayer, 1971) or dentition (Mayer, 1972; Goldschmid, Kotrschal, Krautgartner & Adam, 1980). Although the diet of an individual may change markedly during its life, this is not always the case and there are several examples of species whose diet does not appear to alter over their life history (e.g., Gobiesox meandricus; Johnson, 1970; Halmablennius lineatus: Lee & Chang, 1977). More commonly the diet does change with age and hence size. Size is probably more important in determining diet than age per se because of the change in mouth size as the fish grows. A larger mouth enables a fish not only to eat larger prey, but also to consume a greater range of prey sizes. This positive correlation between prey size and/or diversity and predatory size has been reported in several species (Gobius cobitis: Gibson, 1970a; Spinachia spinachia and others: Kislalioglu & Gibson, 1975, 1976a, 1977; Lepidogobius lepidus: Grossman, 1980; Grossman, Coffin & Moyle, 1980; Parophrys vetulus: Toole, 1980; Oligocottus snyderi, Clinocottus analis, and Artedius lateralis: Yoshiyama, 1980). The functional significance of this relationship is probably that it maximizes the energy intake of an individual relative to the cost of food acquisition. In Spinachia spinachia, for example, the mean size of prey taken corresponds closely to that which minimizes the cost/benefit ratio of prey handling time/prey weight (Kislalioglu & Gibson, 1976a). Grossman (1980) came to a basically similar conclusion for Lepidogobius lepidus, but suggested, in addition, that a switch to large prey by bigger fish may reduce foraging time and hence exposure to predators. Apart from a change in size of prey, the diet often changes in quality as well as quantity (fish eat relatively less as they grow, Sarker et al., 1980; Toole, 1980). Changes in quality may simply be a reflection of prey size, a change from small- to large-bodied amphipods, for example, (Kislalioglu & Gibson, 1977), but there is often a much more drastic change from an initially carnivorous diet to one which may be almost entirely herbivorous. This marked switch in diet with age has been recorded in many species (Boleophthalmus dussumierei: Mutsaddi & Bal, 1969a; Gobius cobitis: Gibson, 1970a; Cebidichthys violaceus: Barton, 1973; Montgomery, 1977; Horn, Murray & Edwards, 1980; Hypleurochilus geminatus: Lindquist & Chandler, 1978; Sarpa salpa: Christensen, 1978b; Coryphoblennius galerita: Fives, 1980a). The change to herbivory is often accompanied by a change in dentition (Christensen, 1978b), and an increase in the relative length of the gut (Barton, 1973; Montgomery, 1977; Christensen, 1978b). There do not seem to be any cases where this dietary change is reversed, i.e., a change from herbivory to carnivory with age, nor

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reports of energetic or foraging studies which would explain the gradual increase in the importance of vegetation in the diet. It may be that the large amounts of protein in the carnivorous diet of young individuals are necessary to sustain their faster growth and higher metabolic rate (Phillips, 1972) whereas carbohydrates, which serve largely as a source of energy (Cowey & Sargent, 1972), can satisfy the requirements of older and slower growing fish. The physiological energetics of feeding behaviour have been studied in only one species, Blennius pholis. This blenny, in common with others (e.g., B. sanguinolentus: Verigena, 1977), has no functional stomach and browses on a wide range of food (Dunne, 1977). Although it can take large single meals (14–9% of body weight in fish of 1–30 g), and can empty the foregut within a few hours (Grove & Crawford, 1980), a single large meal markedly increases oxygen consumption and reduces the respiratory scope for activity (Vahl & Davenport, 1979). The last authors consider that the most efficient feeding method for this species is, therefore, to take frequent small meals, a strategy which the fish adopts in nature. Where the diet of a species has been studied for a year or more, it is also a common finding that feeding intensity and the composition of the food changes annually (Grossman et al., 1980) and seasonally (Mutsaddi & Bal, 1969a; Dunne, 1977, 1978; Fives, 1980a; Grossman et al., 1980; Horn et al., 1980; Summers, 1980). The seasonal change in feeding intensity is usually related to changes in temperature, food intake generally being lower during the colder seasons, but it may also be related to the breeding cycle. Changes in food composition may reflect changes in the availability of food type (e.g., Horn et al., 1980). The subject of food availability and the selection of food types from the great variety of those potentially available has received relatively little attention, although selective feeding is implicit in the fact that fishes can be separated into feeding types such as herbivores or carnivores. A possible exception is the large Chilean clingfish, Sicyases sanguineus, which has been described as a trophic generalist (Paine & Palmer, 1978). In a detailed comparative examination of stomach contents and samples of algae and invertebrates in the habitat Zander (1979b, 1980) and Zander & Bartsch (1972) were able to demonstrate objectively that some food items were preferred over others. Superimposed upon the ontogenetic, annual and seasonal changes in feeding just described, there are also variations on a shorter time scale. Generally speaking intertidal fishes seem to be visual feeders and consequently the feeding activities of most (but not all, see for example Grossman et al., 1980) are mainly limited to the daylight hours. Critchlow (1972) found that all the species he studied in the Gulf of California fed diurnally and Taborsky & Limberger (1980) describe how the feeding rate of Blennius sanguinolentes in the Adriatic gradually increases during the day until dusk when the fish becomes inactive. This change in feeding intensity was considered to be an adaptation which enables the fish to exploit a food resource which slowly increases in nutritive value throughout the day. A remarkably similar feeding pattern has been described for the herbivorous Ophioblennius atlanticus in Barbados (Nursall, 1981). Taborsky & Limberger (1980) also found some evidence that feeding intensity in Blennius sanguinolentus was modified by tidal fluctuations. This direct observation appears to be the only one of its kind for a rocky shore species, although a tidal rhythm of feeding is frequently inferred for fish in this habitat simply because they are emersed or isolated in

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pools at low tide. Experimental verifications of tidally synchronized feeding rhythms in rocky shore species may be lacking because of sampling problems over the high tide period, but where sampling is not difficult, as on sandy beaches, several examples of tidal feeding rhythms have been documented. Most of these examples are to be found among juvenile flatfish which migrate into the intertidal zone with the flooding tide (Tyler, 1971; Gibson, 1973a; Kuipers, 1973, 1977; Wells, Steele & Tyler, 1973; Thijssen, Lever & Lever, 1974; Lockwood, 1980; Summers, 1980). The actual pattern of feeding, however, varies considerably from place to place. Maximum food intake in Pleuronectes platessa may occur at high tide (Kuipers, 1977), on the ebb tide (Thijssen et al., 1974; Lockwood, 1980) or more or less continuously over the whole tidal cycle (Gibson, 1973a, 1980). In other species such as Platichthys flesus (Summers, 1980), and juvenile Pomatoschistus minutus (Healey, 1971b) feeding intensity is greatest at high tide. Adult P. minutus, on the other hand, have maximum stomach contents on a falling tide (Healey, 1971b). On tropical mudflats, mudskippers are also known to feed as the tide ebbs (Kobayashi et al., 1971; Magnus, 1972; Brillet, 1975). The mechanics of the feeding process in one mudskipper, Periophthalmus koelreuteri, has recently been described in detail (Sponder & Lauder, 1981). Variation in the time of feeding is one means of reducing the possibility of competition for food among sympatric species. By so doing species which are sympatric in space become allopatric in time. The best example of this principle is that given by Critchlow (1972) who found that all members of the intertidal fish community he investigated differed in their feeding periods. Furthermore, a species which occurred in two or more habitats altered its feeding times in ways which reduced overlap with other species. Critchlow considered that these differences in periodicity between sympatric species reduced the probability of interspecific encounters. If such encounters were to occur frequently then feeding time would be wasted and feeding efficiency decreased. This phenomenon of change of feeding time from habitat to habitat was taken as evidence that competition for food was taking place. Further evidence to support this contention came from the observation that where there were few species in a habitat their diets were diverse, but where several species were present the diets of each were much more specialized. A convincing experimental demonstration of how the diet of one species can change in the presence of another is given by Edlund & Magnhagen (1981). The gobies Pomatoschistus minutus and P. microps eat similar types and quantities of food (Corophium and Chironomus) when kept separately in aquaria. When kept together, however, P. minutus preferred Chironomus and Pomatoschistus microps fed almost entirely on Corophium. The presence of P. minutus was also found to depress the feeding rate of P. microps. The existence of competition for food in other communities has also been suggested for cottids in California (Yoshiyama, 1980) and for juvenile sparids in South Africa (Christensen, 1978b). In the Californian cottids there was considerable overlap in the type of prey eaten by the three species considered, but competition was reduced by differences in the habitat occupied or the size of prey taken (Yoshiyama, 1980). The South African sparids Diplodus and Sarpa are resident intertidally only as juveniles, the adults are tidal visitors. The juveniles have similar diets, but competition is minimized by separation of the species in time. Diplodus sargus juveniles are present all year round, but those of D.

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cervinus only appear on the shore from October to November and are hence separated from juvenile Sarpa salpa which are found in tide-pools primarily from July to September. At the time when competition between the three species is most likely (July to November), Diplodus sargus changes its diet to include items not taken at other times of year. As the fish grow, competition becomes reduced because the larger juveniles and sub-adults have different diets and feed in different parts of the intertidal zone (Christensen, 1978b). In other communities competition also seems to be minimized by separation of species in space. Kislalioglu & Gibson (1977) found in an investigation of feeding relationships of an assemblage of intertidal and subtidal fishes in a Scottish sea loch, that there were very few large overlaps in diet between species. This dietary separation was caused mainly by each species feeding in a slightly different portion of the habitat or region of the water column. Sasaki & Hattori (1969) considered that there was no competition between the adults of the Japanese gobies Chasmichthys gulosus and C. dolichognathus and that, although there were similarities between the diets of adult C. dolichognathus and young C. gulosus, competition between the two was avoided because they occupied different levels on the shore. A surplus of available food could account for an apparent lack of competition between species, as in the Mediterranean blennies (Zander, 1980) and the cottids Oligocottus maculosus and O. snyderi in British Columbia. O. maculosus and O. snyderi were found to have almost identical diets and the small differences, which were quantitative rather than qualitative, could be explained by differences in feeding behaviour; O. maculosus being less secretive than O. snyderi, took more “open water” organisms such as copepods (Nakamura, 1971). Yoshiyama (1980) also noted that the feeding behaviour of the cottids he studied was related to the type of prey they consumed. Similarly, when the fish in the community investigated by Critchlow (1972) were arranged in a series according to their feeding behaviour in the order grazers-pursuers-ambushers, the percentage of sedentary prey in the diet decreased accordingly. Other descriptions of feeding behaviour are given for gobies by Healey (1971b) and Grossman et al. (1980) and for juvenile plaice Pleuronectes platessa during their intertidal migrations by Gibson (1980). Kislalioglu & Gibson (1976b) investigated the stimuli eliciting prey attack by Spinachia spinachia and found that prey movement and size were the major determining factors. Unusual feeding behaviour in intertidal fishes is described by Herald & Herald (1973) for the amphibious Dialommus fuscus which captures flying insects and by Todd (1976) for Gobionellus sagittula. This goby lives in tide-pools on rocky or muddy shores and is a non-selective feeder. On muddy shores it leaps out of pools at low tide, takes bites out of the mud and then returns to the pools where it washes the mud out of its mouth via the opercular opening and retains food particles on its gill rakers. Larger fish also rake the film of food from the surface of the mud with their teeth. Finally, the effects of intertidal fishes as predators on the fauna and flora of rocky shores are virtually unknown. It may be that because fish population densities are generally small (see p. 367) their effect is negligible compared with invertebrate predators (Paine & Palmer, 1978). There are, however, no published estimates of the biomass consumed on a daily or annual basis which enable this assumption to be tested. Information contained in the descriptions of diets of individual species gives some insight into the large numbers of organisms that can be taken, presumably within one tidal cycle

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(e.g., Fives, 1980a), but such information is sparse because most investigators report diets in terms of percentages. Two other studies, which were basically concerned with the ecology of the prey rather than the predator, are also of relevance in this context. First, Dethier (1980) investigating the ecology of the copopod Tigriopus californiens found that its predator (Oligocottus maculosus) could eliminate copepod populations from tidepools and that a fish could consume over 1000 copepods a day. Secondly, Reimchen (1979) demonstrated that Blennius pholis may be responsible for maintaining colour polymorphism in populations of the gastropod Littorina mariae. Somewhat more quantitative information is available for the effects of predation by intertidal fishes on sandy and muddy shores. Several authors give figures for the numbers or weight of prey consumed by various species of flatfish (Wells et al., 1973; Thijssen et al., 1974; Summers, 1980; Toole, 1980). The most exhaustive study is that by Kuipers (1977) who calculated that young Pleuronectes platessa in the intertidal zone of the Dutch Wadden Sea consumed from 2–4 g dry weight of food/m2 annually. There was no indication that this degree of predation resulted in a food shortage, however, because at the time of maximum predation (April-June), the biomass of the benthos on which the fish fed continued to increase. Healey (1971b) similarly concluded from his calculations of the food requirements of Pomatoschistus minutus that food was not in short supply for this species even though predation on the main food organism (Corophium) was considerable. Contrasting results were obtained with another Pomatoschistus species, P. microps. Berge & Hesthagen (1981) found that the feeding activities of this goby did not alter faunal composition or discernibly reduce the abundance of the benthic infauna in the Oslofjord. Reise (1977), on the other hand, considered that P. microps was largely instrumental in determining the structure and dynamics of the macrofauna of intertidal mudflats in the eastern North Sea.

PREDATION Although predation is frequently assumed to be an important factor controlling the populations of intertidal fishes it has rarely been observed and there have been no detailed studies of its occurrence or estimates of its effects. The proximity of the intertidal zone to land means that it is accessible to both aquatic and terrestrial as well as avian predators, although the general consensus of opinion seems to be that aquatic predators are the most important. Examples of predation by fishes are given by De Martini (1978) on the eggs of Enophrys bison, MacDonald (1975) on Clevelandia ios, and Phillips (1977a) on the embryos of Chasmodes bosquianus. De Martini (1978) considered, however, that birds were unlikely to attack the eggs of Enophrys bison when they were exposed at low tide. Birds as possible predators are also cited by MacDonald (1975), Tasto (1975), Nakamura (1976a), Dunne (1977), and Marsh et al. (1978). Follett & Ainley (1976) found several intertidal fishes in the stomachs of the pigeon guillemot Cepphus columba but thought that these birds were not serious predators on the intertidal fish fauna. Small mammals may also exert a predatory influence (Nakamura, 1976a; Mason & MacDonald, 1980) as may man himself, either directly using the fish as food (e.g., the large clingfish Sicyases sanguineus: Paine & Palmer, 1978) or indirectly

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through localized over-collecting or damage to the habitat (Moring, 1979b). One instance of invertebrate predation is cited by Fives (1980a) in which the actinian Anemonia sulcata captured the small blenny Coryphoblennius galerita. There is also a record of mudskippers being bitten by mosquitos (MacNae, 1968). The main defence of intertidal fishes against predators seems to be their agility, cryptic colouration, visual alertness, and intimate knowledge of their immediate surroundings. This knowledge enables them to gain refuge rapidly when danger threatens. The intertidal area of sandy beaches may be a refuge in itself for many juvenile forms, the shallow water deterring predators from venturing too close inshore (Gibson, 1973; Toole, 1980). Subsidiary defence mechanisms including leaving the water in amphibious forms (e.g., Mnierpes: Graham, 1970b), mimicry of limpets (Zander, 1967), the display of warning signals to other members of the population (Clevelandia: MacDonald, 1975) or taking up defensive postures (Leptocottus: Tasto, 1975). In addition, some species are known to produce toxic mucus (e.g., the clingfish Diademichthys lineatus: Hori et al., 1979) or to contain toxins in other organs. Elam, Fuhrman, Kim & Mosher (1977), for example, found that toxins were present in the gobies Clevelandia ios, Acanthogobius flavimanus, and Gillichthys mirabilis. Their results suggested that the synthesis of the toxins is correlated with reproductive cycle and that it occurs principally in the liver and the ovaries. It is not known, however, whether the toxins are produced primarily as a protection for the adults and/or their eggs or whether they are a by-product of some metabolic reaction. If the latter, then the protective function may be secondary. In this context it is interesting to note that De Martini (1978) suggested that the eggs of Enophrys bison may also be toxic to warm-blooded predators.

CONCLUSIONS Although our knowledge of the biology of intertidal fishes has clearly increased over the past decade, there are still several areas in which much remains to be done. In the author’s opinion the following topics are of greatest interest. (1) The recruitment of larvae to the shore. It is usually assumed that larvae are passively dispersed after hatching. Such an assumption may not be valid, however, and it would be useful to know the distribution of the larvae of resident intertidal species. Do the majority remain close to the shore, and if so, how is this proximity maintained? Alternatively, if they are carried away from the coast, can they return to the intertidal zone to settle? (2) Is larval settlement selective? This seems a particularly fruitful field for further experimental work along the lines initiated by Marliave (1977). Such laboratory studies would need to be complemented by careful observational work in the field to try and detect natural settlement patterns. (3) The phenomenon of zonation is reasonably well documented, but it is a common finding that species overlap markedly in their vertical distributions. Considerably more information is required on the actual microhabitats occupied in the region of overlap as well as the differences in high and low tide distributions. It is quite conceivable that species which are apparently similar in their low tide habitat requirements differ

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markedly in their distribution at high tide. (4) Homing. The evidence for the existence of homing behaviour in certain rocky shore species is convincing, but as yet we have little idea of the mechanisms involved, nor even of the movement patterns of most species at high tide. It would be very useful to be able to track individuals and analyse their movements in detail. There are formidable problems here, however, not the least of which are the small size of the fish, their cryptic habits and the irregular and turbulent nature of the habitat. Perhaps, with the increasing miniaturization of electronic components, a sonic or radio tag small enough to fit these fish will one day be devised. Meanwhile, the alternative will be to choose a species and habitat which allows direct observation of movement and homing behaviour. (5) How are intertidal fish populations regulated? The factors controlling the size and composition of fish populations are virtually unknown, but Grossman (1982) found recently that productivity seemed to be the main environmental factor affecting fish community structure on the Californian coast. The results of repeated tidepool defaunations suggested that, in contrast to those of algae and invertebrates, fish populations were deterministically rather than stochastically regulated. This is an important finding from both a practical and a theoretical viewpoint and further comparable studies would be of value to ascertain whether it is of general application. (6) The rôle of fishes in the ecology of intertidal communities. Although some information is available as to the effect of fishes on the fauna of sandy shores, almost nothing is known of the rôle that fishes resident on rocky shores play as predators. It would be interesting to know whether they exert a significant controlling influence on populations of their prey or whether their relatively small numbers prevent them from having any major impact.

ACKNOWLEDGEMENTS Numerous people assisted in the preparation of this review by supplying references or copies of unpublished theses. Particular thanks in this respect are due to J.B.Graham, M.H.Horn, M.J.Penrith, J.Ruck, D.A. Thomson and P.Wirtz. J.H.S.Blaxter, J.M.Green and M.H.Horn kindly read and commented on all or parts of a draft manuscript. Discussion with participants in the symposium on “Ecology and Physiology of Intertidal Fishes” held at the 61st meeting of the American Society of Ichthyologists and Herpetologists (1981), Oregon State University, also helped to clarify some of my ideas. Finally, I thank my wife for her help throughout.

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ASPECTS OF THE BIOLUMINESCENCE OF FISHES PETER J.HERRING Institute of Oceanographic Sciences, Wormley, Godalming, Surrey GU8 5UB, U.K.

Oceanogr. Mar. Biol. Ann. Rev., 1982, 20, 415–470 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION During the course of the last two decades bioluminescence has been the subject of numerous reviews. The scope of the topic and the extent of the information that is becoming available necessitates that each review should tend to focus upon certain specific aspects and it is no longer appropriate to attempt a general overview. Previous authors in this series (Boden & Kampa, 1964; Tett & Kelly, 1973) have concentrated upon the ecological aspects of plankton luminescence, while a recent multi-author volume (Herring, 1978) has attempted to provide a compendium of information on all aspects of bioluminescence, continuing that achieved by E.Newton Harvey in his classic volume of 1952. The purpose of this article is to use the example of the fishes to focus upon certain ecological aspects of bioluminescence in which substantial advances have been made in recent years. A detailed taxonomic survey of luminescence in fishes has been given by Herring & Morin (1978), which should be consulted for additional information beyond the purview of this paper. Earlier reviews and bibliographies restricted to fishes (Harvey, 1957; Jerzmanska, 1960; Kamykowna, 1960; Nicol, 1967, 1969; Haneda, 1970; Anctil, 1971) have supplemented the data collated by Harvey (1952). The aspects to be considered here are: (1) the involvement of luminous bacterial symbionts, (2) the nature and origin of the luminescent system in those fishes not utilizing bacterial symbionts, (3) the neural control of luminescence, and (4) the special control implied by the concept of ventral counter-illumination. No attempt has been made to impose precise limits on the time scale of the literature under consideration. In general, however, that since 1960 is extensively cited and will provide a route to additional earlier work.

BACTERIAL SYMBIONTS Luminous bacteria are present free-living in sea water, in the gut contents of marine

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organisms and as saprophytes, parasites and symbionts. Free-living luminous bacteria have been obtained from sea water of every ocean region that has been examined. Several species have been characterized and attempts have been made to understand the ecological relationships between those that are free-living and those that occur in symbiotic association. Bacterial symbionts are known to be the source of luminescence in species from at least five different orders of fishes (Greenwood, Rosen, Weitzman & Myers, 1966), the Salmoniformes, Lophiiformes, Gadiformes, Beryciformes and Perciformes (Table I). Additional unpublished data suggests that at least one example also occurs in the Anguilliformes though the bacterial nature of this luminescence is still unconfirmed (Paxton, pers. comm.).

TABLE I Systematic distribution of luminous bacterial symbionts in fishes: *bacteria positively identified by culture and/or electron microscopy; the presence of bacteria is only inferred in the other genera; † the escal bulb is absent in the two ceratioid families Neoceratiidae and Caulophrynidae and in the gigantactinid genus Rhynchactis

Order Beryciformes

Perciformes

Gadiformes

Salmoniformes

Lophiiformes†

Family

Genera

Anomalopidae

Anomalops*, Photoblepharon*, Krytophanaron*

Trachichthyidae

Paratrachichthys*

Monocentridae

Monocentris*, Cleidopus*

Apogonidae

Siphamia*

Acropomatidae

Acropoma*

Leiognathidae

Leiognathus*, Gazza*, Secutor*

Macrouridae

Coelorinchus*, Malacocephalus*, Nezumia*, Sphagemacrurus*, Hymenocephalus, Odontomacrurus, Ventrifossa, Trachonurus, Cetonurus, Lepidorhynchus, Mesobius

Merlucciidae

Steindachneria*

Moridae

Physiculus*, Brosmiculus, Gadella, Tripterophycis

Bathylagidae

Opisthoproctus*, Winteria*, Rhynchohyalus*

Chlorophthalmidae

Chlorophthalmus*

Melanocetidae

Melanocetus*

Oneirodidae

Oneirodes*, Chaenophryne*, and 12 other genera

Diceratiidae

Diceratias, Paroneirodes

Centrophrynidae

Centrophryne, Spiniphryne

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Gigantactinidae

Gigantactis

Himantolophidae

Himantolophus

Ceratiidae

Ceratias*, Cryptopsaras*

Thaumatichthyidae

Thaumatichthys, Lasiognathus

Linophrynidae

Linophryne*, Edriolychnus, Photocorynus, Borophryne, Acentrophryne

SYSTEMATICS AND PROPERTIES OF LUMINOUS MARINE BACTERIA SYSTEMATICS The marine species of luminous bacteria have been the subject of considerable taxonomic revision. Breed & Lessel (1954) proposed that all luminous bacteria should be included in two genera, Photobacterium and Vibrio. Hendrie, Hodgkiss & Shewan (1970) concluded that an additional

TABLE II Nomenclature, temperature properties and natural locations of marine luminous bacteria: modified from Nealson (1978) which should be consulted for the original sources

Designation Hendrie et al., 1970

Reichelt & Baumann, 1973, 1975

Temperature °C Baumann et al., 1980

Growth Storage range

Light Gut organ symbionts l symbionts

Lucibacterium harveyi

Beneckea harveyi

Vibrio harveyi

10–37

18

−

+

—

B.splendida

Vibrio splendidus

10–37

18

–

–

Vibrio fischeri

Photobacterium Vibrio fischeri fischeri

4–25

18

+

+

4–35

18

+

+

P.mandapamensis P.leiognathi

P.leiognathi

P.phosphoreum

P.phosphoreum P.phosphoreum

0–25

4

–

+

—

P.logei2

Vibrio logei

0–25

4

–

–

Alteromonas hanedai1

4–20

—

–

–

1 Data

from Jensen et al., 1980.

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Bang et al., 1978

genus, Lucibacterium, was necessary to accommodate one of the five species to which the known strains could be assigned, all but one (Vibrio albensis) having a requirement of between 0.5 and 3% NaCl. Reichelt & Baumann (1973, 1975) confirmed the single freshwater species Vibrio albensis and five marine species, Photobacterium phosphoreum, P.leiognathi, P. (=Vibrio) fischeri, Beneckea (=Lucibacterium) harveyi, and B.splendida (Table II). Subsequent work has added Photobacterium logei (Bang, Baumann & Nealson, 1978) and Alteromonas hanedai (Jensen et al., 1980) to the marine species. In the most recent reassessment of the known luminous species Baumann, Baumann, Bang & Woolkalis (1980) have proposed the abolition of the genus Beneckea and the assignment of the species in this group, as well as some Photobacterium species, to the genus Vibrio. Since the previous literature on the marine luminous symbionts consistently refers to them as species of Photobacterium this practice will be followed here, notwithstanding the rearrangements noted above. PHYSIOLOGY Details of the many physiological characteristics of the different species of luminous bacteria can be found elsewhere (Nealson & Hastings, 1979) but certain features are particularly relevant to their involvement in symbiotic associations. Oxygen requirement. An analysis of the oxygen requirements for maximum luciferase synthesis and luminescence in species of Photobacterium and Beneckea has shown some interesting differences (Nealson & Hastings, 1977). The rate of synthesis of the bioluminescent system in strains of Photobacterium phosphoreum and P.fischeri is greater at low, growth-limiting, oxygen concentrations than at higher ones. In strains of P.leiognathi and Beneckea harveyi, on the other hand, low oxygen concentration limits both growth and luminescence. Temperature. The optimum temperature for growth varies in different species. B.harveyi and Photobacterium leiognathi will grow at temperatures of up to 35°C but only P.phosphoreum and P.logei will grow at temperatures between 0° and 4° (Table II). Autoinduction. Synthesis of the luminescent system in species of both Beneckea and Photobacterium is greatly enhanced in cultures of high cell densities. This phenomenon is attributed to autoinduction of the system. An inducer is produced by the bacterial cells and accumulates in the medium (Nealson, Platt & Hastings, 1970; Nealson, 1977); when a critical concentration is reached synthesis of the luminescent system is induced. The identity of the inducer molecule for P.fischeri has now been established (Eberhard et al., 1981). In conditions in which inducer cannot accumulate, such as low cell densities maintained by dilution or continuous dialysis of cultures against fresh medium (Ulitzur & Hastings, 1979), the luminescence per cell is greatly diminished. Catabolite repression. Synthesis of the luminescent system in Beneckea harveyi is repressed by glucose, but the repression is overcome by exogenous cyclic adenosine monophosphate (AMP) (Nealson, Eberhard & Hastings, 1972). Photobacterium species behave differently. Strains of P.leiognathi and P.phosphoreum both show glucose

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repression which is not reversed by cyclic AMP while P.fischeri suffers a transient repression when switched to glucose-containing media (Henry & Michelson, 1970; Makemson, 1973; Ruby & Nealson, 1976). Kinetics. The luminescence decay kinetics of the in vitro reaction of bacterial luciferase with reduced flavin mononucleotide (FMNH2) and dodecanal show marked differences between species. Both species of Beneckea have slow decay rates (first order rate constants of 0.11–0.12 s−1) while Photobacterium species have fast kinetics (rate constants of 0.64–1.0 s−1) (Nealson, 1978). The luciferase from the two groups of species can therefore be readily distinguished in vitro. DISTRIBUTION OF FREE-LIVING LUMINOUS BACTERIA The distribution of the free-living marine luminous bacteria has been shown to be correlated with the ambient temperature. In the nearshore surface waters off Southern California, Ruby & Nealson (1978) found that Beneckea harveyi predominated in the summer but was almost absent in winter, whereas Photobacterium fischeri was present all year round and dominant in winter. P.phosphoreum was found only in a few winter samples. In Mediterranean waters Beneckea harveyi was present all year round while Photobacterium fischeri only occurred in winter, and in the Gulf of Elat P. leiognathi was present throughout the year though replaced in coastal surface waters by Beneckea harveyi in the summer (Yetinson & Shilo, 1979). The distribution patterns are believed to be controlled not only by temperature but also by salt tolerance, resistance to photooxidation, and the ability to grow in nutrient-poor conditions (Shilo & Ytenson, 1979). In the estuarine environment of Galveston Island, Texas, only B.harveyi was found, its numbers declining in cold weather (O’Brien & Sizemore, 1979). Samples from Arctic and Antarctic waters consist primarily of Photobacterium phosphoreum and P.logei (Nealson & Hastings, 1979) while P.phosphoreum dominates the midwater Atlantic samples. Below about 1000 m the numbers of luminous bacteria decline markedly (Ruby, Greenberg & Hastings, 1980). Studies of the occurrence and distribution of luminous bacteria in the gut contents of 27 species of surface and midwater fishes have shown that P. fischeri, P.phosphoreum, and Beneckea spp. are often present in concentrations of 105 to 107·ml−1 of gut contents (Ruby & Morin, 1979). In the same study it was demonstrated that a marked species of bacteria when fed to the fish Chromis punctipinnis passed rapidly through the gut and was defaecated in viable form. It is not yet clear whether these enteric bacteria provide any benefit to the fish but it has been suggested that they may aid in the digestion of chitin (Nealson & Hastings, 1979). It is also possible that the luminescence of a faecal pellet might attract a predator, thus recycling the bacteria to other nutrient-rich enteric environments (Robison & Morin; cited by Ruby & Morin, 1979).

DISTRIBUTION OF BACTERIAL LIGHT ORGANS IN FISHES Bacterial symbionts have been identified as the source of light in the light organs of four families of ceratioid anglerfishes (and are probably present in five others) as well as those

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of eleven other families of fishes (Tables I, III). The extent of the information available varies considerably between families and details of the ecology of many species are particularly scanty.

TABLE III Distribution of bacterial species in fish light organs: modified from Nealson (1979); symbols in parentheses are those in which the bacteria have been isolated and characterized but not unequivocally identified.

Bacterial species Family

P. P. P Unidenti Light organ opening leiognathi fischeri phosphoreum fied +1

Anomalopidae Trachichthyidae

(+)

Monocentridae Apogonidae

Rectum

+

External

(+)

Intestine

Acropomatidae Leiognathidae

External

+ +

Rectum Oesophagus

Macrouridae

+

External

Merlucciidae

(+)

Rectum

Moridae

+

Rectum

Bathylagidae

+

Rectum

Chlorophthalmidae

(+)

External +1

‘Ceratioids’ 1

External

No self-luminous cultures have been obtained from these fishes.

BERYCIFORMES Anomalopidae The beryciform family Anomalopidae encompasses three genera and four species, Anomalops katoptron, Photoblepharon palpebratus, Krytophanaron alfredi, and K.harveyi. All four species have a prominent suborbital photophore filled with bacteria (Fig. 1). The photophore of Photoblepharon is relatively larger than that of any other fish. The fine structure in Anomalops has been described by Bassot (1968) and in Photoblepharon by Kessel (1977). That of Krytophanaron is probably similar. The bacteria are contained within a mass of parallel tubules up to 1 mm in length and 30–40

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µm in diameter aligned at right angles to the surface of the organ. Each tubule has a thin epithelial wall and the tubules are arranged in rosettes of five or six around a capillary. Near the surface of the organ the tubules open into collecting channels which coalesce and open to the exterior through numerous pores (Haneda, 1957c) 13–15 µm in diameter in Photoblepharon (Kessel, 1977). The base of each tubule closely abuts on to a reflector which in Anomalops is composed of two parts (Watson, Thurston &

Fig. 1.—a, Anomalops katoptron: b, c, d, Photoblepharon palpebratus: the large bacterial photophore is visible as a white patch under the eye.

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Nicol, 1978). The main interior reflector covers most of the inner surface of the organ. It is composed of stacks of guanine crystals which lie parallel to the surface. Nevertheless because the surface itself is microscopically irregular it has a matt appearance. Its reflectance is ≈70% over most of the visible spectrum (above ≈460nm) and it probably increases the light emission from the organ by ≈35%. A smaller external reflector is present along the lower margin of the photophore. It, too, is composed of guanine crystals, arranged so that the surface acts as a half-silvered specular reflector. It is concluded (Watson et al., 1978) that the effect of this reflector is to increase the proportion of light directed dorsally. The light is controlled either by drawing a black pigmented curtain up from the lower border of the socket of the organ and over its emitting surface (Photoblepharon), by rotating the whole organ downwards into a dark suborbital pocket (Anomalops) or, in Krytophanaron, by a combination of both mechanisms (Rosenblatt & Montgomery, 1976). The musculature for both systems is derived from the levator maxilla. As the bacteria are continuously luminous these movements have the effect of producing a ‘blink’ in the animal’s light. The movements are rapid, up to 100 blinks per minute in Anomalops and Photoblepharon. The family is widely distributed; Anomalops and Photoblepharon occur in the IndoPacific (and the latter genus extends into the Red Sea) while Krytophanaron is a new world genus. K.harveyi has recently been described from Baja California (Rosenblatt & Montgomery, 1976) and K.alfredi has now been reported from several localities in the Caribbean region although before 1977 only the type specimen was known. Divers have reported a group of nine K.alfredi blinking rapidly and swimming together in a volume of ≈1 m3 (Colin, Arneson & Smith-Vaniz, 1979). The photophore was occluded for a brief period (<1 s) when the fish changed direction. The authors consider that the photophores are used to search out prey on the reef face and that the animals normally spend the day in deep water. They are observed at night at depths of ≈10–40 m on reef faces close to deep water (McConnaughey, 1980). Nothing is yet known about the behavioural characteristics of K.harveyi. A considerable amount of information has been amassed on the luminous behaviour of Anomalops and Photoblepharon by Morin and his colleagues in the Red Sea and Indonesia and by McCosker elsewhere in the Indo-Pacific (McCosker & Lagios, 1975; Morin et al., 1975; Doak, 1976; McCosker, 1977; Clark, 1978; Herring & Morin, 1978). Photoblepharon in the Red Sea was first observed in 1964 (Fridman, 1972) and in this area spends the day within caves in the reef, at depths of only a few metres. In the Banda Islands, Indonesia, these fishes occur at depths of >60 m by day. They are extremely photophobic and only emerge to forage on the reef when there is no moon. In Banda, foraging groups are small, up to six individuals, whereas in the Red Sea groups of up to

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50 can be observed. An hour or more before dawn they return to their day-time retreats where they apparently do not feed. During foraging their lights are exposed almost continuously. Fish maintained in constant darkness in the laboratory blink infrequently at night (average values: 2.9 blinks per min, 260 ms duration) but during the day-time there is an almost equal on-off blinking pattern (37 blinks per min, 800 ms duration). High blinking rates are produced when the fish are disturbed and there is a characteristic ‘blink and run’ escape behaviour involving sudden changes of direction during the dark blink period (Morin et al., 1975) The field observations indicate that the light is used to illuminate the prey which may itself be attracted by the luminous aggregations of fishes. The aggregations may be a deterrent to predators and the ‘blink and run’ pattern has a defensive rôle. The observations also indicate that Photoblepharon uses its luminescence in a variety of intraspecific rôles. Observations on Anomalops in the Banda Islands suggest that the general functions of luminescence are similar though the behavioural patterns differ from those of Photoblepharon in the same locality. Groups of up to several hundred roam over the reefs at night blinking constantly (65–70 per min, 250 ms duration) (Haneda & Tsuji, 1971b, 1974). Male-female pairs show distinct sexual patterns. That luminescence is used for feeding purposes is supported by the observation that in captivity dark specimens are relatively unsuccessful at prey capture unless the aquarium is illuminated (McCosker, 1977). When maintained in aquaria the luminescence of individuals tends to decline and finally is extinguished, with a parallel reduction in bacterial numbers within the organ (Meyer-Rochow, 1976) though the reason for this is not understood. Attempts to re-infect dark animals with bacteria from brightly luminous Photoblepharon have failed (McCosker, 1977). A similar gradual dimming has been reported in captive Kryptophanaron alfredi (McConnaughey, 1980). The different behavioural patterns of Anomalops and Photoblepharon, with the shorter illumination periods of the former, are thought to have been derived from a primitive Krytophanaron-type form adapted to a relatively deep habitat (Rosenblatt & Montgomery, 1976). Invasion of shallower habitats has involved modifications of the light organ; the shutter mechanism of Photoblepharon has allowed the photophore to enlarge and to be occluded in short rapid blinks. Its habitat close to the coral substratum renders it difficult for a nocturnal predator to attack despite its bright illumination. Anomalops, on the other hand, having become more of a pelagic planktivore, is more vulnerable to predation but the rapid continuous blinking of individuals in large schools makes it harder for a predator to focus upon a single individual Measurements of the angular distribution (in air) of light emitted by the photophores of freshly killed specimens of Anomalops and Photoblepharon were made in the Banda Islands during the Alpha Helix South East Asia Bioluminescence Expedition in 1975. The apparatus had an angular

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Fig. 2.—The angular distribution of light from the photophores of Photoblepharon (left) and Anomalops (right): the plane in which each pair of measurements were made is indicated by the central diagram (top, horizontal; middle, transverse; bottom, sagittal); in each case the animal was arranged so that the photophore was at the centre of rotation; the relative light output was measured at 10° intervals and is indicated by the length of the line in each direction, the longest line indicates the direction of maximum intensity and has been normalized to the same length in each angular diagram; the angular distribution of light in Photoblepharon is considerably wider than that in Anomalops; in the latter the ventral component is more marked.

acceptance of 12° (Herring, 1976) and the results show marked differences between the two species particularly in the dorso-ventral and horizontal planes (Fig. 2). Photoblepharon with its larger more convex photophore has a considerably broader angular distribution in both planes. Indeed the field of illumination cast by its photophores is remarkably uniform round much of the head. The external reflector in Anomalops (Watson et al., 1978) may perhaps partially compensate for the very limited dorsal illumination otherwise cast by its photophore. Anomalops, however, has some potential for altering the direction of the emitted beam for Morin has noted that it can ‘squint’ its photophores to a limited degree (Herring & Morin, 1978). The nature of the bacteria in anomalopid photophores has not yet been established. Despite numerous early (e.g. Harvey, 1922; Haneda, 1943) and more recent attempts, luminous in vitro cultures have not yet been achieved. Electron micrographs of the bacteria in situ (Haneda & Tsuji, 1971a,b; Kessel, 1977; McCosker, 1977) show them to contain large numbers of storage granules akin to the polyhydroxybutyrate (PHB) granules accumulated by both Photobacterium phosphoreum and P.leiognathi (Nealson, 1978) particularly in nitrogen-limited conditions. Bacterial luciferase has been extracted from the organs (Haneda & Tsuji, 1971a,b) and has recently been demonstrated to have slow kinetics in all three genera (Leisman, Cohn & Nealson, 1980). As these authors point out this leaves open the possibility that the symbionts may be (1) a Beneckea species, (2) a new Photobacterium species with slow kinetics or (3) an entirely new group of luminous bacteria. This problem still awaits resolution. Trachichthyidae Rosenblatt & Montgomery (1976) suggest that the anomalopids and trachichthyids are derived from a common ancestral stock. Of the recent trachichthyids, the five species in the benthopelagic genus Paratrachichthys are the only known luminous forms (Herring & Morin, 1978). The only detailed accounts of the luminous organ are those of Kuwabara (1955) and Haneda (1957a,b). The light organ is a doughnut shaped structure around the anus, and its highly folded epithelial walls enclose numerous bacteria. The light is diffused anteriorly and posteriorly by translucent muscles. There are numerous chromatophores associated with the system; these probably produce the changes in light

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intensity and distribution observed by Haneda (1957a). The bacteria have been cultured in vitro and are probably a strain of Photobacterium phosphoreum. Monocentridae The two known genera (Cleidopus and Monocentris) contain four species all of which are luminous. The light organs are situated one on each side of the lower jaw. Those of Cleidopus gloria-maris have an orange colour which has the effect of altering the spectral emission to a more blue-green colour (Haneda, 1966a). When the fish closes its mouth the upper jaw occludes the light organ, which consists of a mass of epithelial tubes containing bacteria, a reflector, and a superficial lens. The bacteria can readily be cultured (Haneda 1966a; Yoshiba & Haneda. 1967: Yoshiba. 1970: Graham. Paxton & Cho. 1972) and have been identified as Photobacterium fischeri (Fitzgerald, 1977). Graham et al., (1972) noted difficulties in maintaining luminescent subcultures and raised a number of questions concerning the specificity of the symbiosis; these have been largely answered by recent work on Monocentris. The light organs of Monocentris japonicus have been studied by Tebo, Linthicum & Nealson (1979). The light organs take the form of two small protuberances on the lower jaw and lack both a reflector and a lens. The surface of the organ is covered with epidermal papillae; ducts running through four larger papillae open to the exterior. The bacteria are present in numerous tubules which empty into a central channel connected to the exterior through the ducts. The bacteria are closely associated with the surfaces of the cells making up the tubule walls and occasional intracellular bacteria have been observed. The bacteria in situ lack flagella but when grown in liquid medium develop flagella and divide more frequently than appears to be the case within the organ. There is some indication that the tubule cells secrete material into the lumen by exocytosis. Comparison of the bacterial isolates from four specimens of M.japonicus have shown that each animal contained a pure culture of Photobacterium fischeri at densities of up to 94×109 bacteria·ml−1. Many of the physiological features of a representative isolate could be significant in the symbiosis. First, the strain responded to an inducer which might accumulate in the organ; secondly, light production was maximal under low growthlimiting oxygen concentrations; thirdly, the bacteria are not catabolite-repressed by glucose, a probable nutrient present in the blood and, fourthly, the strain excretes pyruvate during aerobic metabolism on glucose. It was suggested that this latter character could be a means of regulatory communication between host and symbiont. The epithelial cells surrounding the bacteria contain two types of mitochondria. These could be responsible for both the lowering of oxygen tension and/or the removal of pyruvate (Tebo et al., 1979). On this basis the Monocentris symbiosis is interpreted as an oxygenlimited system in which the host provides glucose and the bacteria excrete pyruvate (Fig. 3). The pyruvate is taken up by the mitochondria which thus compete with the adjacent bacteria for the available oxygen (Nealson, 1979). PERCIFORMES

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Apogonidae Although many apogonids, from both coastal and oceanic environments, have luminous organs (Herring & Morin, 1978) only one genus, Siphamia, is known to harbour luminous bacteria. The light organ is similar in all the species of the genus and forms a small glandular structure situated dorsal to the ventral musculature just anterior to the pelvic girdle. A number of ducts run from the gland to open into the intestine (Iwai, 1958, 1959, 1971). Light is diffused ventrally by the translucent muscle bands and the emitted intensity can be increased if the fish is disturbed (Haneda, 1965). The eggs and larvae are not luminous (Haneda, 1965) which suggests that bacterial infection is a secondary phenomenon and not transmitted via the egg. The characteristics of the bacteria cultured from the light organ suggest that they are a strain of Photobacterium leiognathi (Haneda, 1966b, 1967; Yoshiba & Haneda, 1967; Nealson, 1979).

Fig. 3.—Models of the bacteria-host association in leiognathids and monocentrids (from Nealson, 1979).

All other apogonids whose luminescence chemistry has been investigated appear to utilize the same luciferin as the ostracod Cypridina (see below). The situation in Siphamia thus represents a curious anomaly. Acropomatidae These fishes are sometimes included in the Apogonidae; there are only two known species, Acropoma hanedai and A.japonicum. The light organ is a U-shaped body straddling the anus, the loop facing in opposite directions in the two species. Numerous ducts from the organ coalesce into a canal which opens externally just anterior to the

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anus. Translucent muscles act as light diffusers and a dorsal reflector and numerous chromatophores serve to control the emission. Although the bacteria have been cultured from both species (and have different characteristics) they have not yet been identified (Haneda, 1950, 1957c). Leiognathidae All 20 species of the three pony fish genera, Gazza, Secular, and Leiognathus, have a doughnut-shaped light organ which opens into the oesophagus. In some species the organ encircles the oesophagus while in others it is displaced or enlarged dorsally (Haneda & Tsuji, 1976) with clear indications of a paired structure. Populations of luminous bacteria are enclosed within tubules lined with a glandular epithelium. The light from the bacteria shines through a chromatophore-occluded window into the swimbladder. This is silvered dorsally and half-silvered ventrally, acting as a parabolic mirror to direct the light into the translucent musculature through which it diffuses over the ventral surface of the fish. The structure of the organ has been described in detail by Ahrens (1965), Bassot (1966, 1975), and Haneda & Tsuji (1976). The last authors examined 18 species and have noted that the males of L. elongatus and L.rivulatus have much enlarged organs, projecting deep into the swimbladder. The light organs of L.hataii and L.aureus are relatively much larger than those of other species (Abe & Haneda, 1972; Haneda & Tsuji, 1972, 1976). Studies of the fine structure (Bassot, 1966, 1975) have shown that the bacteria are invariably extracellular. The epithelial lining of the tubules is rich in endoplasmic reticulum, has a microvillous border and shows signs of secretory activity. Observations of live animals indicate that light is not necessarily emitted continuously but can be stimulated by illumination (Hastings, 1971). The fishes can produce both short bright flashes and steady ventral luminescence (Hastings, 1971; Haneda & Tsuji, 1976; Morin, unpubl.) and it is assumed that muscular control of the aperture of the light organ is involved in the production of flashes while chromatophores exert a longer term intensity control. The bacterial nature of leiognathid photophores has long been recognized and the bacteria have been identified as Photobacterium leiognathi (Boisvert, Chatelain & Bassot, 1967; Hastings & Mitchell, 1971; Reichelt, Nealson & Hastings, 1977). The bacteria from Leiognathus elongatus failed to provide in vitro luminous cultures similar to those of other species (Haneda & Tsuji, 1976) but the significance of this difference is not clear. Bacteria were present at densities of 109−1010·g−1 of organ in four species investigated (Hastings & Mitchell, 1971). Similar bacterial strains were found ‘freeliving’ in the water from which the fish were collected and it was concluded that a single symbiotic strain was involved. More recent investigations (Reichelt et al., 1977) have largely confirmed the earlier work; 21 specimens of six species (from all three genera) were all found to have only Photobacterium leiognathi as their symbionts. The investigation also showed that although most light organs are dominated by a single strain of P.leiognathi some fish definitely harbour more than one strain. Electron micrographs of the in situ bacteria (Haneda & Tsuji, 1971a; Bassot, 1975) give no indication of PHB storage granules, yet when the same bacteria are grown in vitro such granules are present (Bassot, 1975). This suggests that the bacteria within the organ are

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not nitrogen-limited, though they are in certain culture media. The oxygen-limited model for the bacterial symbiosis in Monocentris japonicus (which is probably also applicable to other symbioses involving Photobacterium fischeri and P.phosphoreum) is not applicable to P.leiognathi for this species requires high oxygen tensions for maximum luminescence and does not excrete pyruvate. Nealson (1979) has therefore suggested that leiognathid symbionts are nutrient-limited in continuous culture and that the light organ is effectively a carbon-limited bacteriostat (Fig. 3). GADIFORMES Macrouridae More than half of the approximately 300 species of macrourids have bulbous ventral light organs (Herring & Morin, 1978) which are situated anterior to the anus and from which a duct opens into the perianal groove or the rectum. The morphology of the organ varies considerably between different genera and species. Some are single bulbs, others double structures joined by a long tubule. Lenses and reflector systems may be present and the light may be diffused throughout specialized regions of the ventral musculature. The details of the structure of some of the organs have been described by Haneda (1951) and more recently surveyed (Okamura, 1970; Marshall & Cohen, 1973; Marshal & Iwamoto, 1973). The organ of Mesobius spp. has subsequently been described by Hubbs & Iwamoto (1977). Observations of luminescence in live animals are almost totally lacking. Haneda (1951) observed it in Coelorinchus and Okamura (1970) cites an observation from a submersible. It is likely that the luminescence in some species is brightest in juveniles (Haneda, 1951) or even confined to the young stages (Marshall, 1965) in which the light organ is relatively large (Marshall, 1965; Hubbs & Iwamoto, 1977). Under normal conditions the light is probably regulated by chromatophores. The bulbs of the light organ harbour luminous bacteria which can readily be cultured in vitro (e.g. Haneda, 1951; Ruby & Morin, 1978). Isolates from three species of Nezumia and Sphagemacrurus have been identified as Photobacterium phosphoreum and it is likely that other macrourids harbour the same species of bacterium. Most of the isolates from these three species were referable to a single phenotype. The presence of P.phosphoreum in these deep-water fishes reflects the temperature characteristics of this species (Ruby & Morin, 1978) as well as its better survival at high pressures than P.fischeri or Beneckea harveyi (Brown, Johnson & Marsland, 1942). Merlucciidae Steindachneria argentea is the only known luminescent merlucciid and its luminous system closely parallels that of the macrourid Lepidsorhynchus denticulatus in that it consists of a glandular structure harbouring a dense population of bacteria within its epithelial folds and situated around the rectum into which it opens. Light is similarly emitted through the hypaxial musculature and a characteristic striated area of skin (Cohen, 1964; Haneda, 1968a). Adrenaline injection increased the light intensity by aggregating the chromatophores surrounding the organ (Cohen, 1964; Haneda &

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Yoshiba, 1970). Electron micrographs of the in situ bacteria give no indication of PHB storage granules (Haneda & Yoshiba, 1970); the bacteria can readily be cultured and are probably Photobacterium phosphoreum (Haneda & Yoshiba, 1970). Moridae Luminescence in morids has been observed only in species of Physiculus (Haneda, 1951; Herring & Morin, 1978) but light organs are also present in species of the genera Gadella, Brosmiculus, and Tripterophycis (Marshall & Cohen, 1973). The organs are similar to those of macrourids; a bulbous gland containing bacteria opens into the rectum (Haneda, 1951). Although luminescence has been reported from Lotella phycis (Haneda, 1951) this was a misidentification of some other morid, for Lotella is not luminous (Cohen, 1979). The bacteria from Physiculus rastrelliger have been cultured and identified as Photobacterium phosphoreum (Morin & Nealson, unpubl.). The depth distribution of luminescent species of gadoids is well enough documented to allow Marshall & Cohen (1973) to make some generaliz-ations. They consider that luminescent gadoids are “predominantly slope dwelling fishes of mid-depths ranging from 100–1000 m” and note that no abyssal gadoids are luminous. While most gadoids are benthopelagic in habit there are a number of mesopelagic macrourids including both luminous (e.g. Mesobius, Nezumia) and non-luminous (e.g. Squalogadus, Macruroides) genera. The luminous species do not have depth ranges clearly distinguishable from the non-luminous ones and the rôle played by luminescence in both these and the benthopelagic gadoids remains unknown. SALMONIFORMES Bathylagidae The presence of rectal light organs in the fishes Opisthoproctus, Winteria and Rhynchohyalus was indicated on histological grounds by Bertelsen and his colleagues (Bertelsen, 1958; Bertelsen & Munk, 1964; Bertelsen, Theisen & Munk, 1965). This has been confirmed by observations of fresh individuals and in vitro culture of the bacteria (Herring, 1975). The light organs are rectal diverticula containing clonal populations of Photobacterium phosphoreum which contain PHB granules in situ (Herring, 1975; Ruby & Morin, 1978). The light in Opisthoproctus is directed into a hyaline light guide region which diffuses it over the flattened ventral sole of the fish. A similar but less elaborate system obtains in Winteria, but the anal light organ of Rhynchohyalus probably emits light directly to the exterior as there are no accessory optical structures present (Bertelsen et al., 1965; Herring, 1975). Chlorophthalmidae A perianal light organ is present in Chlorophthalmus albatrossi and C. nigromarginatus (Somiya, 1977, 1981). The organ is structurally similar to that in certain macrourids in that it is a doughnut-shaped body encircling the anus and enclosing a mass of luminous

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bacteria. The bacteria are extracellular though closely associated with the surfaces of the epithelial cells lining the organ and electron micrographs (Somiya, 1981) show storage granules within the bacteria. They can readily be cultured in vitro and are probably referrable to Photobacterium phosphoreum (Somiya, 1981). LOPHIIFORMES Most deep sea angler fishes have a luminous lure in the form of a terminal luminous bulb (the esca) on a modified dorsal fin ray (the illicium). Dahlgren (1928) first suggested that bacteria in the esca were the source of light and recently this identification has been confirmed. Bassot (1966) identified bacteria in the escas of Linophryne and Cryptopsaras (as Mancalias) on the basis of electron micrographs and in Ceratias, Gigantactis and Melanocetus by histological comparison. Hulet & Musil (1968) described bacteria in the esca of Melanocetus murrayi and concluded that they were intracellular. In the esca of Oneirodes acanthias the bacteria are present in deep invaginations of the cells of the tubules within which they lie (O’Day, 1974), and bacteria were also identified in the esca of Linophryne arborifera (Hansen & Herring, 1977). A detailed study of the ultrastructure of the esca of Chaenophryne draco has confirmed the typical appearance and extracellular nature of the bacteria as well as their deep invagination into many of the tubule cells (Munk & Bertelsen, 1980). It is probable that the earlier interpretations of the bacteria as intracellular (Hulet & Musil, 1968) result from the confusing appearance of these invaginations in relatively poorly fixed material. Additional luminous organs are present in Ceratias and Cryptopsaras in the form of sessile bulbs derived from other dorsal fin rays (the caruncles). These also contain dense populations of luminous bacteria (Hansen & Herring, unpubl.). The escal bulb containing the luminous bacteria is a black-pigmented bag with a dorsal or anterodorsal aperture which can be varied, though the mechanism is not clear. The esca also bears a variety of elaborate optical structures including tubular light guides and lenses, so that the light may emerge from a number of separate and widely spaced apertures, some on elongated appendages (Bertelsen & Pietsch 1977; Hansen & Herring, 1977; Munk & Bertelsen, 1980). It is these structures which led Haneda (1968b) to the erroneous conclusion that the luminescence of Himantolophus was intrinsic and not bacterial in origin and to interpret the separate emitting regions as individual light organs. Numerous attempts to culture the bacteria from a variety of ceratioid escas have consistently failed to yield self-luminous cultures (e.g. O’Day, 1974). Nevertheless the presence of bacterial luciferase in the escas of species of Ceratias, Cryptopsaras, Melanocetus, and Oneirodes has recently been demonstrated (Leisman et al., 1980). The luciferase has fast kinetics implying that the symbionts may be a species of Photobacterium. Further evidence for the bacterial nature of the light comes from the fact that the emission spectra from the escas of Oneirodes sp., Chaenophryne and Edriolychnus are almost identical to that of a strain of Photobacterium phosphoreum isolated from Opisthoproctus (Herring, unpubl.).

GENERAL CONSIDERATIONS

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The utilization of symbiotic luminous bacteria as a source of light imposes certain physiological demands upon the fish host. The bacteria need to be restricted to a particular organ without indiscriminate spread throughout the host’s body, the population must be maintained in good physiological condition, excess bacteria must be ejected or absorbed and successive generations of host must be re-inoculated (Herring, 1977a). In all known fish (and squid) symbioses of this type the bacteria are extracellular and this feature may be correlated with the additional characteristic that all such light organs have an opening either directly to the exterior or into the gut lumen. Only in the tunicate Pyrosoma (in which the light organs have no duct) is there any evidence for intracellular luminous bacteria (Mackie & Bone, 1978). One of the most striking differences between fishes with bacterial and intrinsic light organs relates to their number; a single individual may bear many hundred intrinsic photophores whereas one or two bacterial organs is the rule. Although the identification of bacterial luminescence in fishes is now well established and the physiology of the bacteria involved has been extensively studied, the ecological and physiological aspects of the symbioses are largely unresolved. This point was clearly made by Nealson & Hastings (1979) who voiced it succinctly in the question “What are the physiological mechanisms involved in the establishment, maintenance and control of the various symbioses?’ The establishment of the bacterial species from enriched populations in the gut flora is a ready explanation for the source of luminous symbionts in those light organs that open into or near one or other regions of the gut (the great majority). Nevertheless, the data provided by Reichelt et al. (1977) and by Ruby & Morin (1978) show that selection for particular bacterial species is very great in different fishes. Although environmental temperature may be a contributory factor (Ruby & Morin, 1978) the absolute specificity so far identified suggests that a more precise control is maintained over the selection of the bacterial strain. It would be of considerable interest to know whether the specificity in the light organs is mirrored by the gut flora. The data of Ruby & Morin (1979) concerning the relative proportions of different bacterial species in the gut of nonluminous fishes and in the surrounding sea water show, first, that at least three species may be present in the gut flora (though one tends to dominate) and, secondly, that the gut flora does not always reflect the population composition of the surrounding sea water. The latter differences may perhaps be more indicative of the environmental history of the fish rather than its immediate situation. There have not yet been any parallel investigations into the gut flora of fish with bacterial light organs. It is also clear that more than one strain of a bacterial species may be present in one light organ Ruby & Nealson, 1976; Reichelt et al., 1977), suggesting the possibility of multiple infection rather than the single clone condition hitherto assumed. As these authors point out, more detailed investigation of the phenotypic character of the bacteria may show that the apparent dominance of one phenotype is an oversimplification based on insufficient data. It would be most informative to examine in detail the phenotypic variation between the two light organs of single individuals of a species such as Monocentris japonicus now that information is available on the variation to be found among individual fishes. It has also been reported (Hastings, 1975) that antibiotic substances are present in some light organs, which may indicate a mechanism for the biochemical selectivity imposed by the host over and above any physical environmental

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effects. The infection of light organs which do not open into the gut (Anomalopidae, Monocentridae, ceratioids) is more difficult to envisage as a chance event. The very low bacterial densities in the deep sea (Ruby, Greenberg & Hastings, 1980) suggest that some more certain means of infection must be contrived. The current inability to culture some of these symbionts in vitro provides circumstantial support for their more deliberate transfer from one individual to another in vivo. There is as yet no evidence to suggest that the eggs are infected when laid and the possibility of infection by mouth brooding has been suggested for the Anomalopidae (McConnaughey, 1980). Haneda (1965, 1980) also considers this as a possibility in the case of Siphamia in which, although the eggs and newly hatched larvae are not luminous, the larvae are mouth-brooded by the male. If the initial infection is indeed a chance inoculation from environmental sources it is clear that marked specificity of selection occurs within the light organ (Reichelt et al., 1977). Macrourids apparently select for Photobacterium phosphoreum, but only adults have been investigated. Many macrourids have a marked ontogenetic migration and eggs may develop in surface waters (Merrett, 1978) where they may well be exposed to bacterial species additional to P. phosphoreum. Unfortunately there is no information available on the stage at which the light organ becomes functional, but if it were to develop early in life (as in Mesobius berryi), while in surface waters, the possibility of infection by Photobacterium fischeri would be a real one. The question then arises as to whether the symbiont species change during the life history of the fish or whether the selectivity of the photophore milieu overrides that of e.g. environmental temperature (Ruby & Nealson, 1976). If the bacteria of ceratioids are not transferred directly from one generation to another a parallel situation might arise, for larval ceratioids have a much shallower depth distribution than do the adults. The exciting possibility that many of these problems may prove tractable is raised by recent success in keeping certain fishes (particularly anomalopids and leiognathids) in aquaria for extensive periods. This has already allowed Nealson and his colleagues to show that in sterile conditions luminous symbionts are continuously released to the exterior by Monocentris, Cleidopus, and Photoblepharon and that the photophore culture is therefore a continuously growing one. Estimates of bacterial growth rates made from the rates of release indicate a much slower doubling time than in laboratory cultures. The fish thus have some, as yet identified, mechanism for limiting the symbiont growth rate (Nealson, Tebo & Haygood, unpubl). It may soon be possible to rear the larvae of some of these fishes in the laboratory and expose them to different environmental populations of potential symbionts. The effects of such experimental situations on the symbionts that the fish acquire will provide a great deal of information about the mechanisms of symbiont establishment and selectivity. Study of larvae of both coastal and oceanic fish would be valuable to determine the stage at which infection and light organ function become effective. Another practicable approach to an understanding of the specificity of symbiosis is the investigation of bacteria in e.g. leiognathids from as wide a range of geographical regions and water temperatures as possible, as proposed by Reichelt et al. (1977). A parallel experimental study of competition between potential symbionts under different combinations of temperature, pressure, and nutrients would also be very informative.

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The evolution of bacterial light organs appears to have occurred relatively often in shallow water coastal fishes whereas in oceanic conditions intrinsic light organs predominate. It may be that this is related in some way to the relative abundance of luminous bacteria in coastal waters, particularly in the tropics and subtropics. Buck (1978) and Nealson et al. (1980) have considered the evolution of luminous bacterial symbioses in general and exposed many of the limitations of our present knowledge. Only more experimental work can provide the key to the ecological relationships.

CHEMICAL SYSTEMS It has been tacitly assumed in the past that luminous organisms that do not employ bacterial symbionts as the source of light instead synthesize their own luminescent systems. Recent investigations into the biochemistry of a number of fishes suggest that this view is an oversimplification and that some species may have a dietary requirement for at least part of their luminescent system. The development of this concept is relatively recent, and to some extent still contentious. Its development is reviewed here on a chronological basis so that the sequence of events leading to both the introduction of the idea and the abandonment of previous dogmas can be more readily appreciated. It has been known since the latter part of the nineteenth century that non-luminous extracts could be obtained from certain animals which when mixed would produce light (Harvey, 1952). Such systems have been given the generic name of ‘luciferin-luciferase’ reactions; a substrate, ‘luciferin’, is oxidized in the presence of an enzyme ‘luciferase’ in a light-producing reaction. The luciferin is of relatively low molecular weight and heat stable whereas the luciferase is a high molecular weight protein denatured by boiling. The two components are normally extracted separately under different conditions. No chemical identity is implied by the general use of the term luciferin. Indeed all the early work indicated that the luciferase of one species would not react with the luciferin of another unless the two species were very closely related. This led to the belief that different chemical systems had been separately evolved in different groups of organisms, with the implication that the luciferins of different organisms were structurally very different compounds. The general failure of cross-reactions between the systems of unrelated organisms and the consequent belief in the separate identity of the luciferins was derived from the view that the luciferins were highly specific in their activity. It followed from this premise that if luciferins from different organisms reacted with a single luciferase the luciferases must be chemically identical, or very nearly so. In 1958 the first luciferin-luciferase reaction obtained from a fish was reported (Haneda & Johnson, 1958; Johnson & Haneda, 1958). The system was extracted from the pempherid Parapriacanthus beryciformis (now synonymous with P.ransonneti) and the in vitro system was oxygen-dependent. It was not enhanced by co-factors such as ATP and FMN known to be effective with luciferin-luciferase extracts from the firefly and bacteria, respectively. In the same year a luminous apogonid fish Apogon ellioti (=A. marginatus) was described (Iwai & Asano, 1958) and it was shown that not only could a luciferin-luciferase reaction also be obtained from this species but that the system crossreacted with that of the ostracod Vargula (=Cypridina) hilgendorfii* (Haneda, Johnson &

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Sie, 1958). In a further study of this unexpected inter-phylum cross-reaction it was found that although crude extracts of Parapriacanthus did not react with Cypridina material, Apogon extracts reacted with those of both Cypridina and Parapriacanthus (Haneda, 1959; Haneda, Johnson & Sie, 1959). Quantitative data on the Apogon-Cypridina reaction indicated that Apogon luciferase was either less active than Cypridina luciferase or present only in very small amounts in the extracts and that the two luciferases had some differences in their sensitivity * Despite the reassignment of Cypridina hilgendorfii to Vargula hilgendorfii (Poulsen, 1962) the earlier generic name is retained here (in accordance with the literature reviewed) for clarity of comparison.

and optimum pH (Johnson, Haneda & Sie, 1960). Spectroscopic comparisons between the Apogon luminescent system and its cross-reactions with Cypridina material showed that the absorption and emission spectra were virtually identical (Sie, McElroy, Johnson & Haneda, 1961). Similarly crystalline luciferin prepared from Parapriacanthus was indistinguishable from that from Cypridina (Johnson et al., 1961). Luciferin in Parapriacanthus was present in the pyloric caeca as well as in the anal and thoracic light organs. Luciferase, however, was present only in the light organs (Johnson et al., 1961; Haneda & Johnson, 1962a,b). Cypridina, some still luminescing, were found in the stomachs of a minority of individuals. In discussing the possibility that dietary Cypridina luciferin might therefore be the origin of Apogon and Parapriacanthus luciferin Haneda, Johnson & Shimomura (1966) made the point that although the distribution of Cypridina hilgendorfii does not coincide completely with that of the two fishes other species of Cypridina do occur in the area. Although the luciferins were apparently identical the luciferases of Apogon and Cypridina had significant differences in their molecular size, kinetics, and immunological characteristics (Tsuji & Haneda, 1966, 1967). A separate biochemical investigation of the New World species Porichthys porosissimus (Fig. 4) revealed that it too had an extractable luminescent

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Fig. 4.—Distribution of photophores in Porichthys notatus (from Strum, 1969a).

system which cross-reacted with that of Cypridina (Cormier, Crane & Nakano, 1967). Porichthys luciferin was readily extracted from many of the fish’s tissues, though it was most concentrated in the photophores, but luciferase was only obtainable from the photophores, and then with difficulty. The luciferins of Porichthys and Cypridina had similar emission spectra and chromatographic behaviour and were apparently identical. Porichthys porosissimus is an Atlantic species while P.notatus extends up much of the western seaboard of North America. During the course of an investigation into the photophore fine structure of this latter species Strum (1969b) reported the curious fact that specimens from Puget Sound in the north failed to luminesce after adrenaline (epinephrine) injection, whereas those from a more southerly collecting area (Pacific Grove, California) luminesced readily. This apparently trivial observation was to have a seminal influence upon the development of the concept of dietary luciferin. Meanwhile further work on apogonid fishes in the Philippines revealed that a further five species of fish in the genera Archamia, Rhabdamia, and Apogon all gave positive cross-reactions with Cypridina and Parapriacanthus extracts (Haneda, Tsuji & Sugiyama, 1969a,b). The authors noted that Cypridina noctiluca is common in Philippine waters, although none were found in the stomach contents of over 1000 Apogon ellioti that were examined. The work on the Indo-Pacific apogonids and the North American Porichthys was brought together in an extensive comparison of the cross-reactions that could be obtained between them, Cypridina, and Euphausia pacifica, a crustacean present in the stomach contents of many Porichthys specimens (Tsuji, Haneda, Lynch & Sugiyama, 1971).

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Porichthys luciferin produced light with ‘luciferase’ extracts from Apogon, Parapriacanthus, Cypridina, and Euphausia. Euphausia similis luciferase extracts gave a positive result with crude Cypridina luciferin but, significantly, did not react with pure luciferin. A majority of the stomach contents of Porichthys, whether containing euphausiids or only unidentifiable material, reacted with Cypridina luciferase. The crossreactions with euphausiid material were unexpected in that Shimomura & Johnson (1967, 1968) had already shown that the euphausiid system was radically different from that of Cypridina. The specificity of the cross-reactions, and the identity of the luciferins assumed therefrom, appeared much less certain than had hitherto been believed. Kinetic, spectral, and chromatographic data indicated the virtual identity of Cypridina and Porichthys luciferins, and the luciferases of Cypridina, Porichthys, and Parapriacanthus proved to be of similar molecular size (Tsuji et al., 1971). These authors concluded that the evidence suggested that Porichthys may derive its luciferin from ingested euphausiids or other crustaceans. Their data did not rule out a de novo synthesis of luciferin and they pointed out that this could only be resolved by experiments with captive individuals. Attention has subsequently focused on Porichthys, in view of its ready accessibility, ease of maintenance and particularly the fortuitous existence of non-luminous individuals in Puget Sound. These latter individuals contain no detectable luciferin, although luciferase is present in the photophores (Tsuji, Barnes & Case, 1972). Specimens from the Santa Barbara area, on the other hand, contain luciferin in all stages from egg to adult, and even after 150 days of starvation luminesce brightly in the area of a noradrenaline injection. Luciferase is first detectable when the larvae become free-swimming. Tsuji et al., (1972) provided direct experimental evidence of a requirement for exogenous luciferin in Porichthys. One week after intraperitoneal of Cypridina luciferin previously non-luminous Porichthys develop a green fluorescence in the photophores and luminesce in response to noradrenaline injection. Clearly luminescence is induced by the exogenous luciferin. Additional experiments confirmed that the lack of fluorescence and luminescence in Puget Sound fish is correlated with an absence of detectable luciferin (Barnes, Case & Tsuji, 1973). The luminescence intensity of luciferin-injected animals is dose-dependent; luminescence and fluorescence also develop in animals force-fed whole dry Cypridina or saline solutions of luciferin. Luciferin can subsequently be extracted from the fish photophores. An intraperitoneal dose as low as 9 µg, or feeding with a single Cypridina, is sufficient to induce luminescence, albeit at very low intensity. Luminescence can also be induced in non-luminous Porichthys by feeding tissue and skin from intensely fluorescent Californian individuals. Injection of fish with luciferin analogues, oxidized luciferin, or luciferase, does not induce luminescence. Massive noradrenaline injections do deplete the photophore luciferin close to the site of chronic injection, in animals with luminescence induced by feeding very low levels of luciferin (Barnes et al., 1973). The local depletion of photophore luciferin and the dose-intensity relationship suggest that de novo synthesis is unlikely, but do not entirely rule out the possibility of either induced de novo synthesis or the ability to recycle oxidized luciferin within the photophore. Clearly, however, minimal quantities of luciferin will induce luminescence in previously non-luminous animals. Although euphausiids had previously been suggested as a possible source (Tsuji et al., 1972) brightly luminous euphausiids do occur

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in Puget Sound (Barnes et al., 1973). It was concluded that “barring the requirement of some other luminous plankter, the inability of Puget Sound fish to luminesce is not dietarily determined” (Barnes et al., 1973). Cypridina appears to be the “other luminous plankter”. Subsequent data on the emission spectra of the luminescence of induced Porichthys has shown that it is identical to that of Californian animals. Cypridina luciferin has an in vitro emission maximum at 459 nm and extends from 400 to 600 nm. The emission spectrum of Californian Porichthys has two peaks, at 485 and 507 nm, and its green fluorescence is attributable to the known 535 nm fluorescence of Cypridina luciferin. The emission spectra of induced animals also have maxima at 485 and 507 nm (Tsuji et al., 1975). These authors point out that the persistence of induced luminescence far exceeds the known time for which Cypridina luciferin is stable to autoxidation, and note that the delay of 4 to 7 days between dosage and luminescence induction probably reflects the time for transport and concentration of luciferin in the fish photophores, where it is protected from oxidation. What was still not clear was whether normally luminous Porichthys are capable of de novo luciferin synthesis or whether they acquire luciferin from a dietary source not available to the Puget Sound population. Neither Cypridina hilgendorfii nor any other luminous ostracod was known from the range of Porichthys notatus but in 1977 Kornicker & Baker described a new species, Vargula tsujii, whose range overlaps that of the southern population of Porichthys. Many of the remaining problems associated with the Porichthys-Cypridina relationship have been resolved by an extensive study of the zoogeography and luminescence induction in Porichthys notatus (Warner & Case, 1980). This has shown two apparently discontinuous populations, a northerly (non-luminous) one including Puget Sound, and a southerly (luminous) one extending from San Francisco to Baja California. None of the northerly population shows any fluorescence or bioluminescence. The distribution of Vargula tsujii overlaps all but the most northerly part of the southern population. The San Francisco population at this northerly limit includes non-luminous, weakly luminous, and normally luminous individuals. All individuals south of Monterey are strongly fluorescent and luminescent. Both features could be induced by feeding non-luminous fishes with either C.hilgendorfii or V.tsujii, but not by feeding any one of the luminous organisms Gonyaulax, Renilla, Gaussia, Euphausia, Gennadas, Ophiopsila or Stenobrachius. The luminous responses of induced animals to a variety of stimuli are indistinguishable from those of normally luminous individuals, indicating that they have normal control over the induced luminescence. It is possible to deplete the luciferin levels of larvae sufficiently to abolish luminescence. All the evidence thus points to the essential nature of at least some dietary Cypridina luciferin for luminescence in Porichthys and suggests that Vargula tsujii is a normal component of the diet of the luminous southerly population. Information on the chemical nature of the luciferin, such as is now available for some apogonids and Porichthys, is lacking or at best fragmentary for the vast majority of other luminous fishes. Some indications of a peroxidase type of system have been reported in the batfish Dibranchus (Crane, 1968). The situation in the myctophids is somewhat confusing. Initial reports (Tsuji & Haneda, 1971a,b; Haneda, Tsuji, Sugiyama & Hori, 1973) described an oxygen-dependent system extracted from Diaphus which cross-

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reacted with Cypridina extracts. The emission spectra were virtually identical and the reaction between Diaphus luciferase and Cypridina luciferin was inhibited by rabbit antisera to Cypridina luciferase, indicating that the luciferases were similar. Crossreactions between Cypridina material and that from the myctophid Stenobrachius, the hatchetfishes Polyipnus and Sternoptyx, and the decapod shrimp Oplophorus were also noted in these reports. Cormier (1974) obtained cross-reactions between the photichthyid Yarrella and Renilla, the sea pansy. Renilla luciferin has a different structure from that of Cypridina (Cormier, 1978) (Fig. 5) but is the same as that of Oplophorus (Inoue & Kakoi, 1976); this substance has been termed coelenterazine, or coelenterate-type luciferin, and has been isolated and chemically identified from the fish Neoscopelus microchir (Inoue, Okada,

Fig. 5.—Structure of Cypridina luciferin and Renilla luciferin.

Kakoi & Goto, 1977). A luciferase reacting with coelenterazine was also obtained from Diaphus photophores. Coelenterazine has subsequently been extracted from two species of Diaphus, from Yarrella and from the hatchetfish Argyropelecus (Shimomura, Inoue, Johnson & Haneda, 1980) as well as from a number of squid and crustaceans. It seems probable that the earlier reports of Cypridina-type systems in these fishes derived from the relatively non-specific nature of their enzyme systems and that the cross-reactions do not therefore necessarily reflect a Cypridina type of luciferin. The discovery of coelenterazine in Neoscopelus and other fishes, as well as in a number of invertebrates, inevitably leads to speculation concerning a dietary requirement for coelenterazine akin to that for Cypridina luciferin in Porichthys (Inoue et al., 1977). The distribution of coelenterazine is so widespread (Shimomura et al., 1980) that the degree of metabolic conservatism required if all these needs are derived from a single food source renders this concept somewhat improbable. Nevertheless coelenterazine is the basis for luminescence in at least one group of copepods (McCapra, Hart & Herring, unpubl.) and these crustaceans are fundamental components of the food webs in which most larger animals in the marine environment are involved. The overall likelihood of such a general dietary transfer of coelenterate-type luciferin is not high, for it is difficult to envisage the minority of luminous copepods providing the large amounts of luciferin present in such a wide variety of animals. Thus, although there are many examples of animals utilizing chemical compounds of

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dietary origin, as e.g. pheromones and defensive products, it would be premature to abandon the concept of an independent biosynthetic origin for coelenterate-type luciferin in different animals without a great deal more supporting evidence. The most likely situation, on present information, is that different organisms employ different sources of coelenterate-type luciferin, some direct synthesis, others perhaps dietary sources, and a similar diversity of origin probably applies to other luciferins (cf. Dunlap, Hastings & Shimomura, 1980). Isolation purification and chemical identification is the only unequivocal method of determining the nature of an organism’s luciferin. Neoscopelus is the only fish for which this has so far been achieved, and it is impracticable in most cases due to lack of sufficient material. The results of quantified cross-reactions are a second-best approach and, as already demonstrated, need to be interpreted with caution. In many coelenterates the luciferin is bound in the form of a photoprotein (Cormier, 1978) and a luciferinluciferase reaction system cannot be extracted. A similar situation may apply in certain fishes such as the searsiids (Herring, 1972) but the evidence is insubstantial. The chemical difficulties inherent in an alimentary process of luciferin absorption and re-utilization have been noted by McCapra & Hart (1980), and Buck (1978) has discussed the evolutionary implications of such a process. The differences between the luciferases of, for example, Apogon and Cypridina imply that each organism produces its own specific luciferase and that only the luciferin is acquired from a dietary source. The adaptation of the fish luciferase to Cypridina luciferin “would have to be attributed to enzyme induction or to selection from among gene clusters that code for mixed function oxygenases already present in the fish” (Buck, 1978). There is a natural parallel in the selective problems involved in the evolution of complex photophores by fishes which then have to acquire, from the environment, either the luciferin or the bacteria necessary for light production. In order to distinguish more clearly the processes involved in different organisms in general, and fishes in particular, experimental data are necessary on several fronts. One is the identification, as far as is possible, of the luciferin type in different species, even if this is only indicative rather than definitive. Comparison of the luciferases of animals known to use coelenterate-type luciferin would be valuable in establishing their specificity and determining whether each is likely to have been independently derived, as appears to be the case in Apogon and Cypridina. The biosynthesis of both Cypridina and coelenterate-type luciferin is likely to proceed by cyclization of the appropriate tripeptide (McCapra & Manning, 1973). Feeding experiments involving the appropriate labelled amino acids should be attempted; incorporation of the label into the animals’ luciferin would be clear evidence for a biosynthetic rather than dietary origin. Long-term maintenance of captive animals involving deprivation of potential sources of preformed luciferin and, if luminescence is lost, the subsequent feeding with such sources, as in Porichthys, will be difficult for other species but may be possible for some of the coastal apogonids. The minute quantities of luciferin required for luminescence induction in Porichthys make it difficult to distinguish between a continuing dietary source and induced luciferin synthesis. Data on the in vivo turnover of luciferin would help to resolve this point but will be very difficult to achieve. Despite the many experimental hurdles to be cleared, the remarkable story that has gradually unfolded during the course

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of the work on Porichthys and the apogonids provides every incentive for the investigation of other fishes.

NEURAL CONTROL OF LUMINESCENCE Studies on the control mechanisms involved in bioluminescent responses have understandably focused on experimental animals that are more tractable and accessible than the fishes, notably the fireflies, coelenterates and polynoid worms. Information on fishes is in general much less extensive, though Porichthys provides a welcome exception to this rule. General reviews of the physiological control mechanisms of luminescent systems have been provided by Nicol (1955, 1960a), Case & Strause (1978), and (in outline) Anctil (1979c). Despite the limitations of the material, work on fishes has a long history. Some of the earliest observations of neural control mechanisms were made on the photophores of Porichthys, which were shown to luminesce in response to electrical stimulation of the whole animal or injection of adrenaline (Greene 1899; Greene & Greene, 1924). On the basis of these results it was postulated that luminescence is mediated by hormonal control. Adrenaline was later shown to be an effective stimulus for the mesopelagic fishes Echiostoma (Harvey, 1931) and Argyropelecus (Bertelsen & Grontved, 1949). PORICHTHYS The midshipman has continued to be a subject for experimental investigation, particularly in North America, as a consequence of its ready availability and ease of maintenance in aquarium conditions. Nicol (1957) showed that electrical stimulation of the posterior part of a severed spinal cord caused a luminescent response to spread anteriorly after a latency of several seconds, provided the sympathetic nerve chain was left intact. Ligaturing the heart, to prevent any blood-borne hormonal stimulus, failed

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Fig. 6.—Diagrammatic representation of the sympathetic outflow of a teleost fish including suggested innervation of the integumental photophores and chromatophores (from Nicol, 1967).

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to affect the response. Nevertheless injection of adrenaline into an un-ligatured heart will produce a general glow, though with a latency of up to two minutes. Nicol (1957) therefore proposed that control was mediated via the longitudinal sympathetic chain. Postganglionic fibres run from the ganglia via the grey rami (Fig. 6) to join the cranial and spinal nerves (Nicol, 1967). General luminescence can also be induced by direct stimulation of the

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Fig. 7.—a, diagram of a Porichthys photocyte with its associated supportive cell and nerve (after Strum 1969a); b, diagram of a Porichthys photophore showing the relative positions of the different elements

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(from Herring & Morin, 1978).

brain, with the lowest thresholds being found at three sites in the medulla oblongata (Demski & Stepien, 1978). A detailed ultrastructural study of the photophores of Porichthys was subsequently undertaken by Strum (1966, 1968, 1969a, b). She observed many synaptic vesicles in the endings of the predominantly unmyelinated nerves adjacent to the supportive, photogenic and lens cells, but none of the endings was seen to penetrate the basal lamina and make direct intercellular contact with either a supportive or photogenic cell (Fig. 7). The prolonged latency to electrical stimulation was considered to reflect the time a putative neurotransmitter would take to diffuse across the extracellular channel between the nerve termination and the photocytes. Nicol (1967) noted that the sensitivity to adrenaline probably indicated that the post-ganglionic fibres were adrenergic, releasing adrenaline “or more likely noradrenaline” at the photocytes. Noradrenaline and amphetamine both induce glowing; although eserine and acetylcholine do also, the delay with these is too long for a direct neuro-photocyte effect to be likely (Case et al., 1970). Subsequent work on Porichthys has utilized the electrical and pharmacological responses of isolated photophores, in conjunction with developmental studies, to elucidate some of the mechanisms involved in neuro-photocyte control. Baguet & Case (1971) found a progressive deterioration in the condition of isolated photophores in saline, evidenced by an initial increased sensitivity to electrical stimuli and culminating in a spontaneous glow. Single stimuli delivered to spontaneously glowing photophores produced a transitory reduction in light output but after the glow they became completely unresponsive to any stimulus. Freshly isolated photophores responded with latencies of 100–200 ms to single pulses delivered through microelectrodes (Fig. 8). Externally applied stimuli were ineffective at frequencies of less than 4–5·s−1, but multiple short (7 ms) stimuli at 10·s−1 produced summation of the light responses. Repetitive trains of stimuli showed facilitation and a progressive reduction in the latency of the response to each train. Both the rate of light emission and the total light emitted were reduced by lowering the pH from 7.4 to 5.6. Light output was increased in 100% oxygen and prevented by prolonged exposure to nitrogen. Spontaneous light emission followed the re-admission of oxygen after stimulation in anoxic conditions. Maximum light emission occurred at 20°C with a Q10 of 4.5 between 10° and 20°C. Baguet & Case (1971) suggest that as the total number of stimuli delivered in the latent period is similar at rates of 5, 10, and 20·s−1 the effect of each stimulus is additive, culminating in a threshold for light production. Baguet (1975) noted that fast and slow flashes, differing only in their time to peak intensity, not in latency or decay kinetics, could be obtained from different groups of photophores. Light responses to single stimuli were

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Fig. 8.—Responses to stimulation of isolated Porichthys photophores: a, b, and c, P.myriaster (from Baguet, 1975); d and e, P.notatus (from Anctil, 1979b); a, responses to (A) a 10s train of stimuli (12 s−1, 0.5 ms, 20 V) and (B) adrenaline 10−5M; b, responses of a photophore evoked at 60s intervals by 5 s trains of stimuli (12 s−1, 0.5 ms, 12 V); c, time course of flashes evoked by stimuli (150 ms, 100 V) at intervals of 5s (A) and 10s (B); d, light emission evoked by application of 2, 4dinitrophenol (DNP) at 10−4M: e. 10−4M KCN.

proportional to both the strength and duration of the stimulus. Summation and potentiation of responses followed low frequency stimuli (Fig. 8). When two stimuli were delivered to a photophore at short intervals (<10 s), the second response was of greater intensity and faster decay than the first. This was interpreted as indicating that extinction is not a simple depletion of intracellular substrate. High frequency stimulation with trains of long-duration stimuli caused spontaneous glowing and the photophore became unresponsive. This is regarded as indicative of damage, similar to the ageing process. Trains of short duration stimuli (2 ms) did not induce such glowing. The latency of the responses to repetitive stimuli did not alter, in contrast to the previous findings with repeated trains of stimuli (Baguet & Case, 1971). This is believed to reflect differences in the site of action, single stimuli directly affecting the photocytes while trains of stimuli act on neural elements presumably via a transmitter substance. The pharmacological responses of isolated photophores are complex. Photophores are sensitive to both adrenaline and noradrenaline (Fig. 8). Adrenaline produces luminescence at concentrations down to 10−9 M (Baguet, 1975); noradrenaline is also effective at these concentrations but the responses to it are very oxygen-dependent and only some 10% of the magnitude of those to adrenaline at similar concentrations. Both drugs produce an initial rapid flash but adrenaline also induces a long lasting glow (Christophe & Baguet, 1981). At the same time as light emission is stimulated oxygen consumption is decreased (Baguet, pers. comm.). A refractory period to further adrenaline stimulation of about 30 min ensues, during which the photophore still responds to electrical stimuli. The alpha-adrenergic agonist phenylephrine induces a rapid flash while the beta-adrenergic agonist isoproterenol induces a long lasting luminescence. Phentolamine (an alpha-adrenergic antagonist) blocks specifically the rapid response to adrenaline and noradrenaline, while propanolol (a beta-adrenergic antagonist) irreversibly inhibits both the long-lasting response to adrenaline and the response to electrical stimulation (Christophe & Baguet, 1981). The alpha-antagonists yohimbine and ergotamine irreversibly reduce the responses to both electrical and chemical stimuli. Dopamine reduces the response to adrenaline but has no effect on electrical stimulation. Acetylcholine is without effect. 5-hydroxytryptamine (5-HT) reduces the response to adrenaline but can reversibly inhibit electrical sensitivity. In the light of these results it has been concluded (Baguet, 1975; Christophe & Baguet, 1981) that adrenaline is the likely neurotransmitter for the long-lasting excitatory response, noradrenaline for the rapid flash response and that 5-HT might act as the neuromediator for a putative inhibitory innervation. The earlier electron microscope studies of neurovesicles (Strum, 1969b) are interpreted as providing support for this hypothesis.

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Recent work on the effects and sites of uptake of catcholamines and 5-HT (Anctil, Brunel & Descarries, 1981 ; Gariépy & Anctil, in press; Anctil, pers. comm.) has shown that noradrenaline and 5-HT, but not adrenaline, are taken up by the photophores. Noradrenaline accumulates in the axon terminals while 5-HT accumulates within the photocytes themselves. Uptake and storage of noradrenaline is unaffected by 5-HT but photocyte accumulation of 5-HT is markedly reduced by noradrenaline. 5-HT depression of adrenaline-induced luminescence could not be demonstrated (cf. Baguet, 1975). 5-HT inhibits the luminous response to repetitive pulse trains separated by long delays but does not affect facilitating pulse trains separated by short delays. Anctil and his colleagues consider these results point to a presynaptic modulatory role for 5-HT, the 5-HT perhaps being released from the photocytes, and do not believe it necessary to invoke an inhibitory innervation. Other bioluminescent systems with excitatory innervation show spontaneous light emission in raised K+ media. In Porichthys increased levels of K+ and Na+ reversibly decrease the light intensity of isolated photophores while increased Ca2+ has the opposite effect (Baguet, 1975). These results are interpreted as indicating that the ionic effects operate on an inhibitory rather than excitatory system. Conflicting results, however, have been obtained by Martin & Anctil (pers. comm.) who found that K+ induces luminescence and depolarizes the photocytes. Baguet’s (1975, 1977) conclusion that innervation is basically inhibitory and that the development of spontaneous glowing reflects the failure or poisoning of this inhibitory system (in saline, high Ca2+, low Na+ or K+, potassium cyanide (KCN) or 2, 4-dinitrophenol (DNP) Fig. 8) also assumes that the photophore would otherwise luminesce in response to enhanced blood levels of adrenaline. Blanching of the fish is assumed to be a sympathetic response, presumably involving increased adrenaline levels, yet it is not invariably associated with luminescence. Further evidence consistent with this inhibitory hypothesis is that photophores removed from fish in a highly excitable state cannot be stimulated, while those removed from the same fish at other times respond normally. Heightened activity of the postulated inhibitory pathway is assumed to be associated with the increased excitability. An alternative approach to the resolution of the details of the neurotransmitter control was employed by Anctil & Case (1976). They studied the effects of 6-hydroxydopamine (6-OHDA), a specific agent for chemical sympathectomy which destroys noradrenergic and dopaminergic nerve endings almost exclusively. Subcutaneous injections of 6-OHDA induced a luminous response, similar to that of noradrenaline. Low dosages irreversibly abolished photophore responses to electrical stimuli over a 24 h period, leaving a residual glow. These animals showed an initial increased sensitivity to noradrenaline but later also became unresponsive to this stimulus. Fishes given large doses of 6-OHDA showed little or no response to noradrenaline on any subsequent occasion. Electron microscopy of the photophores at different stages after 6-OHDA injection demonstrated a progressive destruction of nerve endings during the supersensitivity phase and increasing numbers of damaged photocytes, supportive and even lens cells as the response to noradrenaline diminished and finally disappeared. The authors concluded that 6-OHDA initially impairs neuro-photocyte transmission by destroying catecholaminergic nerve endings (Anctil & Case, 1976). During this initial phase the endogenous transmitter is released, producing

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the observed low-level glow and the increased sensitivity to exogenous noradrenaline. The subsequent reduction and loss of noradrenaline sensitivity is correlated with a gradual degeneration of the photocytes. The cause of this degeneration is not clear but the chemical sympathectomy may remove some trophic effect of the innervation (Anctil & Case, 1976). The observed responses to 6-OHDA can therefore be explained without involving an inhibitory innervation. An ultrastructural study of photophore development (Anctil, 1977) showed that there is an initial ‘aneural phase’ in which they become fluorescent but luminesce only when treated with hydrogen peroxide. In the subsequent or ‘neural’ phase the photophores also respond to electrical and pharmacological stimuli and become densely innervated; the nerve terminals have narrow neurophotocyte gaps and are filled with large numbers of light-cored vesicles (Fig. 9). The onset of fluorescence and luminescence capability is correlated with the presence in the photocytes of large vesicles filled with a dense flocculent matrix, similar to that observed in the adults (Strum, 1969a). The fluorescence is inversely related to the light output of the photophore when stimulated, declines most rapidly during bright light emission, and is reduced without light emission on treatment with oxidative agents (Baguet, 1977; Zietz-Nicholas & Baguet, 1978; Baguet & ZietzNicolas, 1979). It is concluded that the fluorescence is a reflection of the content of reduced luciferin within the photocytes. The ultrastructural changes occurring in the photocytes following electrical, chemical, and pharmacological stimulation provide additional evidence for the control mechanism (Anctil, 1979a,b). Stimulation of the spinal cord induces changes first in the nerve profiles and then in the photocytes, in which the vesicles coalesce, their contents become aggregated and the microvilli are reduced in number. Noradrenaline stimulation has a similar effect upon the photocytes but no marked effects on the nerve terminals. These morphological changes are reversible, intimating that they are unlikely to be artifacts. Specialized synaptic contacts, with neuro-photocyte gaps of ≈50 nm, probably represent the sites of transmitter action. Treatment of fishes with KCN and DNP produces glows (see Fig. 8) which are correlated with alterations in the nerve endings and these precede changes in the photocyte morphology. Blockage of neurotransmission by alphaadrenoceptor antagonists and reserpine abolished the luminous response to DNP. The results indicated that DNP and KCN act first and foremost on the neural processes of Porichthys photophores. Both compounds are mitochondrial inhibitors and mitochondria actively take up calcium. The effects of DNP and KCN can therefore be interpreted in terms of the liberation of sequestered mitochondrial calcium; calcium in turn is known to be the trigger for transmitter release at most peripheral neuroeffector junctions. A model for intracellular control of luminescence based upon calcium regulation by the mitochondria has recently been put forward (Zietz-Nicolas, Thines-Sempoux, Zurstrassen & Baguet, 1980). Despite the impressive correlation between luminescence, photophore development and photophore ultrastructure there is still no clear consensus over the presence of an inhibitory innervation. The initial interpretation of short latencies as direct stimulation of the photocytes (Baguet & Case, 1971) has become less certain with the demonstration by Warner & Case (1980) that response latencies of whole fishes to a variety of stimuli may be as low as 200–300 ms, similar to those achieved by electrical stimulation. It has also

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been convincingly argued that the effective concentration of adrenaline at the site of transmitter action may be very different to that in the blood and that a general sympathetic discharge need not, therefore, trigger luminescence (Case & Strause, 1978). The effects of 6-OHDA support the concept of an

Fig. 9.—Development of luminous responses and correlated structural features of the photophores of Porichthys notatus (modified from Anctil, 1977).

excitatory catecholaminergic innervation yet Baguet’s (1975) data can be interpreted as evidence for an inhibitory innervation. The pharmacological results still do not distinguish unequivocally between alpha- and beta-adrenoceptor activity, though a dominant alpha activity has been reported (Anctil & Gariépy, cited by Anctil, 1979c). Although the early concepts of blood-borne adrenaline as the neuroendocrine basis for the control of luminescence (Greene & Greene, 1924) have largely been abandoned in principle, the details of adrenergic control in practice still await resolution. LANTERNFISHES (MYCTOPHIDAE) The lanternfishes are among the most abundant and certainly most accessible of oceanic fishes. Many species can be captured at the surface at night and they are therefore considerably more accessible as experimental material than are most other meso- and bathypelagic fishes. All species (except Taaningichthys paurolychnus) have serial photophores, largely ventrally and ventrolaterally placed and many species have in addition supra- and/or infracaudal organs, larger blocks of luminous tissue which may show sexual dimorphism (Fig. 10). Early accounts of myctophid luminescence (see Harvey, 1952; O’Day, 1972) described brief flashes or glows produced spontaneously or in response to a variety of stimuli. Subsequent reports have described bursts of flashes or glows, erratic responses to

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electrical stimuli, a lack of any response to adrenaline or 5-HT injection, and a general luminescence on treatment with hydrogen peroxide solutions (Nicol, 1958; Clarke, Conover, David & Nicol, 1962; Anctil & Gruchy, 1970; Anctil, 1972; Baguet, 1975). The detailed anatomical relationships of the nerve supply to the photophores (e.g., Ray, 1950; Anadon, 1957; Lawry, 1973b) and the preliminary experimental results all indicated that the photophores of myctophids were probably under neural control. This was demonstrated elegantly and conclusively by Barnes & Case (1974) who were able to elicit repeated and predictable luminous responses from the body and caudal photophores of seven species of myctophid. Spontaneous displays in undisturbed animals usually involve only the body photophores, whereas the caudal organs flash rapidly in response to some disturbance. The intensity of the body photophores can be rapidly and synchronously varied. Caudal organ flashes occur as a series of discrete high-intensity rapid flashes in the series of subunits making up each organ. Infra- and supracaudal organ responses do not necessarily occur simultaneously and most spontaneous caudal organ flashes had a duration of 60–80 ms. Electrical stimuli (5 ms pulses, 100·s−1, delivered in trains of 0.5·1.0 s) to the spinal region readily triggered whole animal responses rising gradually in intensity and decaying rapidly at the end of the stimulus. Individual photophores of the head and tail responded within 50 ms of each other. Analysis of caudal organ responses showed consistent facilitation and latencies of 25–35 ms unaffected by stimulus strength or duration or by temperature changes in the range 10°–18°C. Stimuli were followed at frequencies of up to 20·s−1 (Fig. 10). A double flash was produced only when paired stimuli were separated by at least 40 ms (≡25 stimuli·s−1). The

Fig. 10.—The lanternfish Ceratoscopelus townsendi showing the position of

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the serial photophores, luminous patches (1,3) and the caudal organs (2): inset are the responses fo electrical stimuli at respectively 10, 20, and 30·s−1 of three such regions in C.townsendi (after Barnes & Case, 1974).

interflash period during spontaneous flashing (≈120 ms) corresponds to ≈8 flashes·s−1. The responses of other patches of luminous tissue in Ceratoscopelus townsendi were similar to those of the caudal organs. Injection of fishes with noradrenaline, eserine, acetylcholine, tubocurarine or gallamine had no significant effect on their luminous responses. Studies of the fine structure of the photophores and caudal organs have shown a fundamental similarity in each though their morphological arrangement differs between species (O’Day, 1972; Edwards & Herring, 1977). The detailed neuroanatomy of the photophores has been examined (O’Day, 1972; Anctil & Case, 1977) and both spinal and sympathetic fibres are involved in caudal organ innervation. Multiple nerve branches are sandwiched between the stacks of thin lamellate photocytes with neurophotocyte gaps of 10–15 nm. Gap junctions are present between adjacent photocytes and, if they electrically interconnect all the photocytes within a photophore or caudal organ unit, could account for the rapid and simultaneous displays (O’Day, 1972; Edwards & Herring, 1977; Anctil & Case, 1977). Some anatomical evidence for adrenergic neurotransmission is given by Anctil & Case (1977) though there is as yet no pharmacological support for this and they consider electrotonic junctions equally probable. In a comparative study of the photophore structure and function in 14 myctophid genera Edwards & Herring (1977) described similar photophore and caudal organ responses to mechanical stimuli, with high frequency (up to 20·s−1) fluctuations in overall photophore intensity and up to tenfold increases on stimulation, with waves of intensification sometimes running from head to tail. Luminescent responses have been elicited from isolated photophores of Diaphus holti (Baguet, 1975; Baguet & Marechal, 1976). The flashes in response to stimuli of 4 ms duration were very short (5–10 ms), reached peak intensity in <1 ms and had latencies as short as 1.4 ms. A stimulus of 8–16 ms duration produced 2 or 3 flashes. The very short latencies suggest that the flash responses do not reflect normal neurophotocyte transmission but rather direct electrical stimulation of the photocytes themselves (Case & Strause, 1978) or even excitation of local nerves and electrotonic transmission at neurophotocyte junctions (Anctil, pers. comm.). The data concerning the control of myctophid luminescence are still very limited. The presence in the photophores of both myelinated and unmyelinated nerve fibres, probably of spinal and sympathetic origin respectively, suggests that neural control is more finely tuned and complex than the simple excitatory pathway so far experimentally demonstrated. The question of neurotransmitters or electrotonic transmission (or a combination of the two) remains unresolved. Of the pharmacological agents investigated so far none appears to have any central or peripheral effect and the apparent immutability of the latency to electrical stimuli (Barnes & Case, 1974) is also in contrast to related phenomena in other systems. Clearly one of the main aims in future work must be the

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isolation of a functional neuro-photocyte unit as an experimental preparation, akin to that achieved in the case of Porichthys. The fragility of these fishes makes this a particularly challenging goal; nevertheless indications that repetitive responses can be obtained from Myctophum photophores (Baguet, Christophe & Marechal, 1980; Christophe & Baguet, 1982) suggest that it may be near to attainment. OTHER FISHES Data concerning the putative neural control mechanisms of other fishes are extremely sparse. Harvey (1952) provides references to earlier observations including his own (Harvey, 1931) demonstration that injection of adrenaline caused the serial photophores of the stomiatoid Echiostoma to glow steadily and that the cheek (postorbital) organ flashed on handling. Similar responses to adrenaline have been reported in Argyropelecus (Bertelsen & Grontved, 1949), Maurolicus (Anctil, 1972) Stomias, Chauliodus, Idiacanthus (Denton, Gilpin-Brown & Wright, 1972), Gonostoma and a number of stomiatoid genera (cited by Herring & Morin, 1978). Rapid flashing or bursts of flashes have been observed in many stomiatoid fishes (Harvey, 1931; Filimonov & Chumakova 1969; Herring & Morin, 1978) but the detailed control systems have been investigated (by Baguet and his colleagues) in only a few species. In general, the responses of isolated photophores of fishes such as Argyropelecus, Ichthyococcus and Chauliodus respond differently to long duration (>4 ms) high intensity (>20 V) stimuli and to short (1.4 ms) weak (<15 V) stimuli (Baguet, 1975; Baguet & Marechal, 1974, 1976; Christophe, Baguet & Marechal, 1979; Baguet et al., 1980; Christophe & Baguet, 1980). The response of Argyropelecus hemigymnus photophores to long, strong stimuli is one or more fast flashes with latencies of <6 ms and total durations <10 ms. Although a distinction is made between the rapid fatigue of Argyropelecus photophores and the greater number of flashes obtainable from those of Chauliodus (Baguet et al., 1980) the differences are primarily of degree. The responses to short, weak stimuli are quite different: at frequencies of at least 1·s−1 they initiate a slow response with a long latency and gradual intensity increase (Fig. 11). The latency decreases and intensity increases with stimulus frequencies up to 20·s−1 in Argyropelecus. In Chauliodus repeated short stimuli produce a sustained emission by summation of individual flashes while in Argyropelecus the response is achieved by potentiation of subthreshold stimuli. The results for Chauliodus indicate that the duration of individual flashes increases as the stimulus duration and intensity is reduced (Christophe et al., 1979). The extinction kinetics of these fused responses vary but are always relatively slow (>10 s). The initial decay is exponential but in Argyropelecus a glow may be initiated which is superimposed upon it and may even exceed the intensity achieved during stimulation (Baguet et al., 1980) (Fig. 11). A glow of this type may even be induced by short weak stimuli without any prior luminous response. The glow is reversibly quenched by superimposed stimuli, in proportion to the strength and duration of the stimulus (Fig. 11). The initial green fluorescence of Argyropelecus photophores decreases in intensity during the luminescent responses. The isolated photophores of Chauliodus and Stomias failed in early experiments to respond to 10−6–10−3 M adrenaline, noradrenaline, acetylcholine or 5-HT. The intensity

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of photophores already feebly glowing was nevertheless increased by 10−5 M adrenaline (Baguet & Marechal, 1976). More recent results (Baguet, pers. comm.) have shown that Chauliodus photophores isolated from fish in better condition do respond to adrenaline, noradrenaline, phenylephrine, and isoproterenol at 10−4 M. Detailed studies of the pharmacology of Argyropelecus photophores (Baguet & Marechal,

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Fig. 11.—Responses to stimulation of isolated photophores of Argyropelecus

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hemigymnus (a-e) and Chauliodus sloani (f): a, fast response, electrical stimuli of 25V, 8ms at 1·s−1; b, slow response to 450s stimulation (10V, 1ms, 4·s−1); c, afterglow following stimulation for 5s (10V, 1 ms, 10·s−1); d, glowing inhibition induced by four trains of stimuli (35V, 100·s−1) with stimulus duration of successively 1, 2, 3, and 4ms; e, dose-response curves of isolated photophores to adrenaline, note the two different populations of photophores (continuous and broken lines), ×, rear photophores, ●, front photophores: f, series of flashes evoked by stimuli (65V, 8ms) at 2·s−1 and 1·s−1; a-d from Baguet et al. (1980; e from Baguet & Marechal, (1978); f from Christophe et al. (1979).

1978) have shown that those that are feebly glowing are maximally stimulated by 10−5– 10−6 M adrenaline. Dose-response curves show some stimulation remaining at 10−9 M and inhibition at 10−3 M, with a separate population of photophores with maximum response at 10−7–10−8 M (Fig. 11). The most effective concentration of noradrenaline is 10−4 M but its effectiveness is only about 0.5% that of adrenaline. The alpha inhibitor phentolamine at 10−5 M depresses the response to adrenaline and blocks that to noradrenaline. The beta inhibitor propanolol by itself stimulates light production in isolated photophores and acts as a competitive inhibitor of the adrenaline responses. It is concluded from these results that adrenaline is a more likely neurotransmitter than noradrenaline, largely on the basis of the photophore’s greater sensitivity to it. Excitatory control is thus envisaged as a low threshold liberation of an adrenaline neurotransmitter. Disturbance of this excitatory system causes spontaneous glowing (perhaps by neurotransmitter release following damage to the nerve endings). The inhibition of this glow by electrical stimuli is explicable either as a direct effect upon the photocyte membranes or as the result of an inhibitory control system. Such a system could be of higher threshold than the excitatory system and either involve a different neurotransmitter or produce a sufficiently high concentration of adrenaline to be inhibitory (Baguet et al., 1980). The differences observed in the experimental data for Porichthys, Argyropelecus, and Chauliodus are largely of degree. The classification of responses in terms of light output and stimulus frequency that Baguet and his colleagues have achieved may reflect more the condition of the preparations than fundamental differences in their neurophysiology. Certainly a comparison between the features of Porichthys, Argyropelecus and Chauliodus indicates some remarkable similarities which would not be expected considering the taxonomic disparity between them. Fast responses probably represent direct electrical stimulation of the photocytes (Case & Strause, 1978) whereas the slow responses are more representative of normal physiological control. Baguet (pers. comm.) considers that the fundamental difference between Porichthys and Argyropelecus photophores lies in the control system at photocyte level, and is reflected by the inhibitory effect of KCN and DNP on adrenaline-stimulated luminescence of Argyropelecus photophores. Unfortunately little is known of the in vivo nature of the light emission of Argyropelecus and Chauliodus. The features of freshly caught specimens (Denton et al., 1972; Herring, 1977b) suggest a steady luminescence of the ventral photophores, while in

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Chauliodus and other stomiatoids the postorbital and certain other organs flash rapidly (Herring & Morin, 1978). Electrical stimulation of the postorbital organs induces flashes whose temporal characteristics are closely akin to those of spontaneous flashes (Herring, unpubl.). It would be particularly valuable to compare the experimental responses of the ventral photophores and postorbital organ of a stomiatoid such as Chauliodus, knowing that their normal responses are quite different. Both can be stimulated by adrenaline in fresh specimens. Much more is known about the normal luminescence of lanternfishes and, as a result, the experimental work of Barnes & Case (1974) can be appreciated as an effective analysis of the normal responses. In physiological terms the stimulus-response relationship of the myctophids (Barnes & Case, 1974) may be compared to the slow responses of Argyropelecus and Chauliodus. Despite their general similarities no pharmacological effects of adrenaline or other agents have yet been demonstrated in myctophid photophores and this system may indeed differ significantly from others if electrotonic junctions control the coupling of photocytes within the photophore.

COUNTER-ILLUMINATION The serial photophores of a majority of fishes (as well as of many crustaceans and cephalopods) are predominantly ventral in their distribution. Even those present on the flanks are generally ventrally directed. Dahlgren (1916), when considering cephalopods, first noted that “any animal situated below such a squid and looking upward at it from below would see a bluish light that would blend with the sunlight”. The same concept was applied to fishes by Rauther (1927) who remarked that this ventral bioluminescence “könnte ihnen eine Angleichung der Bauchseite an die von oben her einfallenden bläulichen Reste des Sonnenlichtes als Schutz gegen Sicht von unten von Wert sein”. Although reiterated by Jerzmanska (1960) and Fraser (1962) little heed was taken of the hypothesis until it was developed and extended by Clarke (1963), though without reference to the earlier literature. The hypothesis holds that animals living at depths still subject to significant light penetrating from the surface will, when viewed from below, be silhouetted against this background illumination; this tell-tale silhouette is reduced or eliminated by means of ventrally directed bioluminescence. This ‘counter-illumination’ camouflage closely parallels the rôle of ventral countershading in surface waters and other well-illuminated environments. Recent considerations of the concept have been provided by several authors (Herring, 1977a; Buck, 1978; Marshall, 1979). This attractively simple hypothesis is founded upon circumstantial anatomical evidence and in recent years considerable effort and ingenuity has gone into the acquisition of experimental and observational data that may aid in determining its general validity. The evidence to date falls into two main categories, first the relationship between the size and distribution of ventral photophores and the ambient light régime to which the animal is exposed, and secondly the degree to which the emitted bioluminescence matches the physical features of the background light. PHOTOPHORE DISTRIBUTION AND AMBIENT LIGHT

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The distributional data supports a counter-illumination rôle for the serial ventral photophores of many fishes; the depth-related differences in photophore size and distribution have long been recognized and documented (e.g. Brauer, 1908; Murray & Hjort, 1912) and will be dealt with here only in outline. In surface waters day-time light intensities are too great for them to be matched by bioluminescence: camouflage is achieved by adaptations of body shape and surface reflectivity (including countershading) (Denton,

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Fig. 12.—Angular and spectral distribution of light in the sea: a, changes in

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angular distribution with increasing depth (Jerlov & Fukuda, 1960; after Denton, 1970); b, diagram showing the distribution of radiance in the sea, the upwelling component is exaggerated for the purposes of illustration, in the ocean the ratio of upwelling to downwelling light is ≈1:200 (from Denton, 1970); c, change in spectral distribution of downwelling light with increasing depth (from Boden, Kampa & Snodgrass, 1960).

1970). In the upper mesopelagic realm day-time intensities are attenuated sufficiently to be matched by bioluminescence (Fig. 12). The background is, however, still relatively bright and the contrast provided by a silhouette would be very marked and hence easily distinguished. In the lower meso-and bathypelagic realms ambient light intensities progressively diminish until they fall below the threshold level for visual detection of a silhouette (at depths of 700–900 m in oceanic water). At still greater depths (the oceans have an average depth of 3800 m) no significant light penetrates from above and such light as there is derives from in situ bioluminescence. At all depths below about 200 m the radiance distribution of downwelling light is symmetrical about the vertical and remains constant regardless of increasing depth (see Denton, 1970, for discussion) with the upward component only some 0.5% of the downward component. Similarly the spectral distribution of downwelling daylight becomes more and more nearly monochromatic with increasing depth (Fig. 12) and has a transmission peak at about 475 nm in clear ocean water (Kampa, 1970). The photophore distribution of fishes, and other animals, can be correlated with these depth-determined changes in the quality of the background light (Marshall, 1979). Fish which are day-time inhabitants of near-surface waters lack photophores. Those in the upper mesopelagic realm, where ventral counter-illumination is likely to be of greatest value, have numerous large photophores whose distribution is largely restricted to the ventral surface, typified by the hatchetfishes and many gonostomatids and photichthyids (e.g. Maurolicus, Valenciennellus, Ichthyococcus). Fishes from deeper and darker waters have ventral photophores which are relatively much smaller and more widely spaced. In addition, numerous other photophores may be present elsewhere in the body. In species restricted to depths below the penetration of surface light the serial ventral photophores are most often either lacking or are greatly regressed in the adults. The differences in the degree of development of ventral photophores is most clearly apparent in closely related species with different daytime depth distributions. In the genus Cyclothone, for example, the deep species C. microdon and C.acclinidens have much smaller photophores than the shallower species C.alba, and the deepest species C.obscura has only rudimentary photophores. Similarly the deep species Gonostoma bathyphilum has very much smaller photophores than the shallower G.elongatum and G.atlanticum. Similar gradations in photophore size can be found in many other groups of fishes. Many fishes undergo considerable depth changes during their development, with the larvae living in relatively shallow waters and the adults much deeper. The photophore development in these species provide a parallel in that deep water adults with small photophores develop from shallower-water larvae and juveniles in which the photophores are relatively much larger. In addition the photophore development matches the increasing opacity of the animal and photophores tend to develop first beneath the heavily pigmented regions of the body

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(Badcock, 1977). Even in such a fish as Xenodermichthys copei, whose adult distribution extends into deep water and which has photophores distributed all over its body, the juveniles, which have a shallower distribution, develop the ventral photophores first (Badcock & Larcombe, 1980). Similar features apply to other groups of animals with ventral photophores (euphausiids, decapods, and squid).

PHYSICAL CORRELATES OF VENTRAL BIOLUMINESCENCE However compelling such anatomical correlations may appear to be they are in themselves insufficient for a confirmation of the counter-illumination hypothesis. If the animals are to match their background perfectly then they must emit light whose angular distribution, spectral distribution, and intensity is very similar, if not identical, to that of the background light. The angular distribution of the light emitted by the ventral photophores of Argyropelecus affinis and Chauliodus sloani has been determined by Denton et al., (1972). The angular distribution of that from Argyropelecus affinis in both the transverse and longitudinal planes almost exactly matches that of the background radiance in the sea (Fig. 13). The structural basis for this radiance distribution in Argyropelecus photophores is the orientation of reflective guanine crystals on the inner and outer margins of the photophore tube (Denton, Gilpin-Brown & Roberts, 1969; Denton, 1970). Multiple reflexions between the silvered inner surface and half-silvered outer surface determine the emitted radiance distribution (Fig. 13). This is clearly demonstrated by the observed differences in radiance distribution following removal of the outer half-silvered surfaces of a luminescing photophore group (Herring, 1977b). The angular distribution from Chauliodus sloani is similar to that of Argyropelecus, but a little broader in the transverse plane with relatively more light emitted laterally. It therefore matches the background distribution less well. The ecological significance, however, lies not in the measured bioluminescence radiance distribution so much as in the perceived radiance distribution. This is the sum of the bioluminescence and the background light reflected by the flanks of the fish. The flanks of Argyropelecus are so effectively silvered that they act as mirrors, and provide near-perfect camouflage for most of the fish simply by virtue of their reflective capabilities (Denton, 1970). The ventral bioluminescence is solely concerned with camouflaging the narrow silhouette of the ventral margin. In Chauliodus the flanks are less silvered and more rounded. The greater lateral component of the bioluminescence distribution compensates, at least in part, for the reduced reflectance of the blue ambient light from its flanks and will improve its camouflage in ventro-lateral view. Radiance distributions similar to that of the background light have also been measured from the ventral photophores of some decapods and euphausiids, which may also have a counter-illumination rôle (Herring, 1976; Herring & Locket, 1978). For counter-illumination to be an effective tactic it is also necessary for the spectral quality of the emitted light to match that of the background. The only exception to this would be the hypothetical situation in which the predators spectral sensitivity is completely uniform over the whole visible range. Bioluminescence from counterillumination systems would therefore be expected to have an emission maximum around

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475 nm and a relatively narrow bandwidth. Unfortunately there are relatively few measurements of the emission spectra of serial ventral photophores of fishes, but more information is available for Argyropelecus than any other fish. The light emitted by this fish traverses a cellular plug in the aperture of the photophore. The plug contains a reddish purple pigment which has a maximum transmission at about 480 nm (Denton, Gilpin-Brown & Wright,

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Fig. 13.—Photophore structure and function in Argyropelecus spp. (from Denton, 1970): a, Argyropelecus aculeatus; b, diagrammatic view of the anal group of photophores showing the common photogenic chamber (PH) the external half-silvered reflector (R) and the internal argenteum (A); c, section through a single reflecting tube (T) which is filled with a viscous material, the pigmented filter (F) lies in the aperture of the photogenic chamber; d, diagram to show the way the internal argenteum spreads the light over an arc anteriorly and posteriorly; e, the multiple reflexions that would affect a pencil of light projected into the reflecting tube from above; f, the angular distribution of light in the transverse plane of Chauliodus sloani (×) and Argyropelecus affinis (●) compared with a computed curve for daylight in the sea (solid line) and in Lake Pend Oreille (dotted line) (from Denton et al., 1972).

1970; Herring, 1977a). It will therefore act as an effective colour filter more closely matching the spectral distribution of the emitted light to that of the background light. Measurements of the emission spectra of Argyropelecus photophores before and after the removal of this filter have confirmed this conclusion (Denton & Herring, 1978, and unpubl.). Filter pigments are present in the photophores of many other meso- and bathypelagic fishes; the absorption spectra of the pigments extracted from three other species (Herring, 1977a), also have a transmission window in the 470–490 nm region. Most bioluminescence spectra extend further into the longer wavelengths than does ambient light in the sea and these pigments will have the effect of producing a narrow bandwidth spectrum very similar to that of background light. In other fishes (e.g. myctophids such as Diaphus rafinesquii) the spectrally selective properties of the photophore are determined by the reflectance characteristics of the reflector behind the photogenic tissue (Denton & Herring, 1978). The spectral distribution of light from the lanternfish Myctophum (Nicol, 1960b) and Gonichthys (Herring, unpubl.) is very close to that of ambient daylight, with peaks in the 470 nm region. Other spectra are less easy to fit into a counter-illumination context; that of Gonostoma atlanticum has a peak at 507 nm which does not accord with the expected spectral match (Swift, Biggley & Napora, 1977). The significance of this is not clear. The close correlation between the spectral emission of Argyropelecus photophores and that of ambient light is shared by the bioluminescence of several euphausiid and decapod crustaceans (Herring & Locket, 1978; Herring, unpubl.) and certain cephalopods (Young & Mencher, 1980). The last authors have shown that some squid can vary the spectral distribution of their emitted light and suggest that this is an adaptation to enable the animal to match both the spectral radiance of daylight in deep water and that of moonlight nearer the surface. At mesopelagic depths the spectral and angular distributions of ambient light will vary little with depth, though in near surface waters the variability may be more marked. The intensity of ambient light, however, changes markedly with depth. Even in clear oceanic water the light intensity is reduced by at least 96% for every 100 m increase in depth (Clarke & Denton, 1962). If a fish is able to emit light only at a fixed intensity it will achieve effective counter-illumination only at the depth of that particular isolume and its

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vertical movements will necessarily be determined by the daily excursions of the isolume. While there is evidence that the depth of some scattering layers is related to that of particular isolumes during part of their diurnal movements (Clarke, 1966; Boden & Kampa, 1967) this is not a frequent occurrence and the distributions of most species are not so closely determined. It is therefore to be expected that counter-illuminating species are able to control the intensity of emission to match that of their preferred environment and maintain their camouflage relatively independent of fluctuations in ambient light intensity. Considerable effort has been expended in attempts to determine whether or not oceanic animals are able to regulate the intensity of emission in relation to ambient light levels, and to establish what physiological mechanisms may be involved. Many animals luminesce in response to light stimuli (see Herring, 1978, for references) but observed responses of potential counter-illuminating systems are relatively few. Hastings (1971) noted that leiognathids responded to illumination by luminescing; subsequent observations (Morin et al., cited by Herring & Morin, 1978) have confirmed that the luminescence intensity is proportional to the ambient light level. There is also some circumstantial correlation between the size of leiognathid light organs (or their reported intensity) and the light conditions in which the species are normally found (Pauly, 1977). The serial photophores of the lanternfish Tarletonbeania crenularis were reported (Lawry, 1973a, 1974) to luminesce with “a light…which approximated the colour and intensity of the incident illumination” when the fishes were lit from above with a blue light. Further experiments on the lanternfish Symbolophorus californiensis (Case, Warner, Barnes & Lowenstine, 1977) provided strong support for a counter-illumination function by showing that unrestrained fishes varied their luminescence in relation to changes in ambient light intensity over the range of 2−75×10−3 µW·cm−2, similar to those that the fishes would be likely to experience in their normal environment. Recent work on Myctophum obtusirostrum (Young et al., 1980) has provided detailed evidence of the species’ ability to match ambient intensities over a range exceeding three log units and including the intensities to be expected at its normal day-time depth. Rapid changes in ambient intensity were matched by the experimental animal in a matter of seconds (Fig. 14). The only other report of a counter-illumination capability for a fish is for the ceratioid Cryptopsaras couesi (Young & Roper, 1977) which has no known photophores (other than the esca) but which luminesced from the general skin surface. More experimental data are available for invertebrates; those for cephalopods, in particular, are as detailed and at least as compelling as for myctophid fishes (Young & Roper, 1976, 1977; Young et al., 1980) while that for certain decapod crustaceans is almost as extensive (Warner, Latz & Case, 1978; Young et al., 1980; Warner, 1981). What feedback system enables the fish to regulate its luminescence intensity so appropriately? Many species with well-developed ventral photophores also have a preorbital photophore apparently directed into the eye. Nicol (1967) first suggested that this might serve as a reference photophore which could be matched against the observed ambient light; certainly the preorbital and ventral photophores of Argyropelecus do wax and wane in concert (Herring, 1977b). By no means all fishes with serial ventral photophores have specific preorbital organs, but there are often one

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Fig. 14.—a, bioluminescent responses of Myctophum obtusirostrum to changes in overhead light intensity; b, counter-illumination measurements of M. obtusirostrum related to the depth equivalent to the experimental light intensities, ▲ indicate one specimen, ● and are two different runs on a second specimen; c, effects of exposing the pineal organ (left) and eyes (right) of M.spinosum to the overhead light, in each case the luminescence intensity increases rapidly, upward arrows indicate removal of the opaque shield covering the respective organs, downward arrows its replacement, 0 indicates when the overhead light is turned off; a and b, from Young et al. (1980); c, from Young et al. (1979).

or more others directed into the eye (e.g. Tchernavin, 1953). Thus the optical anatomy of the myctophid Tarletonbeania is consistent with the supra-orbital photophore having the rôle of a reference source (Lawry, 1973a,c, 1974). Occlusion experiments have demonstrated that shading either the eyes or the pineal organ of Myctophum spinosum from the overhead light causes the fish to reduce the intensity of its ventral bioluminescence (Young, Roper & Walters, 1979) indicating that both organs are involved in the sensing of ambient light intensities, though the eyes clearly predominate (Fig. 14). It is perhaps significant in this context that McNulty & Nafpaktitis (1977) and McNulty (1979) have suggested, on the basis of the comparative anatomy of the pineal complex of seven species of myctophid, that the deepest living species, Parvilux ingens, has a much more light-sensitive pineal than the shallower Tarletonbeania crenularis, and the deep Cyclothone acclinidens a more sensitive pineal than the shallower C.signata. Work on cephalopods (Young, 1973, 1977, 1978; Young et al., 1979) and the shrimp Sergestes similis (Warner et al., 1978) has indicated that in the former the extraocular photosensitive vesicles play a greater rôle in the regulation of counter-illumination than do the eyes, whereas in the latter the eyes are the prime determinants of background intensities. The experimental evidence to support the hypothesis of a counter-illumination rôle for the ventral photophores of many fishes (and other animals) is already extensive and, when allied to the more indirect inferences drawn from anatomical and distributional data, becomes compelling. Nevertheless in view of the formidable technical difficulties of monitoring (without interfering with) the behaviour of individual fishes in the mesopelagic environment, it is unlikely that conclusive evidence, which in situ observations of luminescence and predation could provide, is likely to be achieved in the near future. Work of this nature could more easily be undertaken on coastal species, such as the leiognathids, but a demonstration of the validity of the hypothesis in the midwater environment is required for its complete corroboration. As Buck (1978) so clearly recognized “refinements of measurement or demonstrations of correlation are no substitute for direct observation of protection in establishing the validity of the camouflage idea”. Whatever the possible mechanisms involved in a feedback control of bioluminescence intensity “an emitting fish has no criterion for the quality of this match but the hard pragmatic test of being able to reproduce before being eaten” (Buck, 1978). McAllister (1967) has raised a number of objections to counter-illumination as a

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hypothesis to account for all cases of ventral positioning of photophores but clearly much more information is required on the general biology of many species with ventral photophores before the justification of considering their position independently of their function can be established. Some of the difficulties inherent in a day-time concept of counter-illumination (e.g. in near benthic species) may be less of a stumbling block if a night-time camouflage rôle is additionally considered. Young et al. (1980) have discussed this situation in oceanic species and come to the intriguing conclusion from their data that some species at least are present on moonlit nights at depths where the light intensities are higher than those at which counter-illumination is likely to be effective. From the prey’s point of view, however, any reduction in silhouette contrast may be better than none. It is generally assumed that counter-illumination implies a continuous luminescence by the fish but observers in submersibles have reported quiescent fishes randomly oriented at their day-time depths and not luminescing (e.g. Barham, 1971). Although Beebe (1935) described the steady glow of hatchetfishes swimming past the bathysphere’s port, it may well be that ventral illumination is normally only initiated when the fish is alerted by some other stimulus to the presence of a potential predator. If the fish is not luminescing continuously the maintenance of a horizontal posture is less critical and, as Marshall (1979) has pointed out, the more nearly vertical fishes (such as lanternfish) hang in the water, the less silhouette will they present (cf. Backus et al., 1968). Certain anatomical features of some meso- and bathypelagic animals have been interpreted as adaptations to ‘break’ the effectiveness of ventral counter-illumination in the prey. Muntz (1976) argues that the yellow lenses in the tubular eyes of Scopelarchus analis serve such a purpose. The yellow pigment acts as a cut-off filter for wavelengths of light shorter than about 450 nm. Muntz assumes that though the ventral bioluminescence has a peak wavelength similar to that of downwelling daylight it has a greater bandwidth, in particular extending further into the longer wavelengths beyond 500 nm. The lens pigment will reduce the amount of ambient light reaching the retina by almost two thirds but will reduce the bioluminescent light rather less because of its greater long wavelength component. Thus although the overall sensitivity will be considerably reduced Muntz (1976) calculates that the contrast of the prey will be increased from 1:1 (perfect camouflage) to 1:1.8. The relative effectiveness of such a system is, of course, also dependent upon the absorption characteristics of the retinal pigment. Data concerning the emission spectra of ventral photophores are still too limited to assess the general validity of the bandwidth assumption (though it can hardly apply to those, such as Argyropelecus, with appropriate filter pigments) but it will certainly be applicable in cases such as Gonostoma (Swift et al., 1977). Yellow lenses are also present in certain species of Argyropelecus (Somiya, 1976), Chlorophthalmus (Somiya, 1977), Echiostoma (Somiya, 1979), Malacosteus (Herring, unpubl.), and perhaps Diretmus (Munk, 1980), as well as the squid Histioteuthis (Muntz, 1976), but are not widely distributed. Another adaptation that may serve to ‘break’ the effectiveness of ventral counterillumination is the convexicliveate fovea of notosudid fishes (Steenstrup & Munk,. 1980). This interpretation assumes that the angular distribution of bioluminescence perceived by the notosudid eye differs slightly from that of the background light. Experimental

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measurements of radiance distribution have employed apparatus with a much greater acceptance angle than that of the notosudid eye, so such small angular differences cannot be ruled out. In any event even if the angular distribution is normally identical to that of the background light slight changes in the orientation of the luminescing prey will produce detectable differences in the radiance distributions. Steenstrup & Munk have calculated that in such a fovea the changes in off-centre foveal illumination derived from the moving image of the bioluminescent (point) source should theoretically be distinguishable from the static background light. No doubt if ventral counter-illumination is as widely employed a tactic as appears probable at present, evolutionary pressures will have led to the development of other visual adaptations to combat its effectiveness.

CONCLUDING REMARKS Consideration of these relatively limited aspects of the bioluminescence of fishes serves more to outline the inadequacies of our present information than to highlight the data that have been acquired. Inevitably, too, focusing on the fishes tends to give a very distorted view of the rôle of bioluminescence in the marine environment. It is one of the most widespread and characteristic phenomena in the oceanic realm and its significance in this particular ecosystem is undoubtedly still greatly underestimated. Nevertheless it has been appropriate to concentrate upon the fishes for they have a greater diversity of known luminous systems (and therefore functions?) than any other single group of animals. Reviewing four particular topics within this wide diversity has made it possible to relate aspects of their anatomy, biochemistry, neurophysiology, and ecology to their bioluminescence. It is evident that the earlier anatomical emphasis is now being leavened with experimental and biochemical studies, with the result that it is becoming possible in certain cases to transform hitherto purely descriptive work into investigative exercises aimed at verifying (or falsifying) particular hypotheses. Most of these hypotheses necessarily still concern capabilities and physiological mechanisms rather than functions. Their development into functional aspects requires more observational data on luminescent behaviour. Buck (1978), albeit in a wider context, stated that “obviously all of the questions cited above, whether of fact or interpretation, require more data. The need for intensive experimental and field studies of living luminous animals is urgent”. The full potential of such studies is almost impossible to appreciate. Fireflies have relatively few photophores of rather similar type: their behavioural complexity is only now being realized. Fishes have a very great range of photophore structure and numbers; contemplation of the likely extent of their behavioural repertoire is at present as frustrating as it is exciting.

ACKNOWLEDGEMENTS I am most grateful to the many people who have, variously, commented on the manuscript, allowed me to quote their unpublished data or discussed the concepts

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involved, in particular M.Anctil, J.R.Badcock, F.Baguet, J.F. Case, P.M.David, N.R.Merrett, and K.H.Nealson. Mrs C.Darter prepared the figures and Mrs P.Talbot typed the various drafts of the manuscript.

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THE BIOLOGICAL IMPORTANCE OF COPPER IN OCEANS AND ESTUARIES A.G.LEWIS Department of Oceanography, University of British Columbia, Vancouver, British Columbia, Canada and W.R.CAVE Workers’ Compensation Board of British Columbia, Vancouver, British Columbia, Canada

Oceanogr. Mar. Biol. Ann. Rev., 1982, 20, 471–695 Margaret Barnes, Ed. Aberdeen University Press

INTRODUCTION Copper is essential for normal metabolism in most organisms but is toxic at elevated levels and forms the principal ingredient of many antifouling compounds. It also enters into many reactions with naturally occurring and introduced organics. These reactions change the chemical state of the metal and its availability to organisms. Copper also interacts with particulates such as clays and colloids which change the state of the metal and hence its biological availability. Once within the organism, the effect of copper may also be controlled by physiological and biochemical factors, some of which vary from species to species. Copper enters marine and estuarine waters from run-off, aeolian input, and geothermal sources. Of these, run-off is probably the most important, providing localized input into coastal waters that are major centres of population and sources of seafood. Processes occurring during and after the entry of copper into the marine environment can cause a change in metal speculation with a resultant change in the biological effect and levels of accumulation of the metal. These processes form what could be called the marine “biogeochemical” cycle of copper. The possibility of release of metals, including copper, through deep-sea mining, disturbance of sediment from nearshore dredging, mine tailing discharge, and wastewater input has caused some concern about potential metal hazards. The biological availability of copper is, however, controlled not only by concentration but also by metal speciation and thus many of the processes involved in the biogeochemical cycle. Biological effects as well as copper levels in organisms are due to the uptake of the metal and its interaction with metabolic processes. As the copper levels in organisms

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may, in part, be due to the levels in the environment, these levels must form part of any review of the biological importance of the metal. Copper levels in organisms are also discussed as are levels of copper found to be toxic to marine organisms. Although a discussion of metal analysis techniques is important to an understanding of the meaning of metal levels in the environment or in organisms, it could form the basis of a literature review by itself. In addition, the ramifications of specimen treatment and the problems of sampling and contamination add another dimension to any discussion of measurement techniques. For these reasons, as well as the desire to restrict discussion to the central topic of the biological importance of copper, it was decided not to include a techniques section. The reader is referred to the excellent reviews by Schmidt (1978a, b) for a discussion of techniques of measuring copper in the environment and in tissues. The present review developed from a series of literature reviews on the biological importance of copper in the sea, written for the International Copper Research Association, Incorporated (INCRA). Through the assistance of Dr Charles H.Moore and Dr George Cypher of the New York INCRA office and Mr Brian Moreton of the London office it has been possible to produce a more complete review. The review has also benefited from the able assistance of Mrs Louise Parker and from the staff of the Woodward Biomedical Library at the University of British Columbia. As a final note, both of us recognize that it is far easier to review the original research of others than to do original research. We only hope that, within this review, we have been able to generate some new ideas or perhaps polymerize existing material in order to produce ideas that will warrant future investigation. To assist in this the literature review draws upon work done in fresh water and soils as well as the marine environment.

THE EFFECTS OF COPPER ON ORGANISMS General reviews of the literature concerning the biological effects of copper are found in a number of publications (e.g., Elvehjem, 1935; Lamb, Bentley & Beattie, 1958; Engle & Woods, 1960; Bowen, 1966; O’Kelley, 1974; Leland, Copenhaver & Wilkes, 1975; Leland, Wilkes & Copenhaver, 1976; Leland, Luoma, Elder & Wilkes, 1978; Schmidt, 1978a,b; Prosi, 1979). For ease of discussion this section of the review is divided into two parts, one on the beneficial effects of copper, the other on the detrimental effects. BENEFICIAL EFFECTS Copper is one of the elements which have been shown to be essential for life processes (e.g., Bowen, 1966) although required in only trace amounts (i.e., <0.1% of the total atoms present). It is present in more than a dozen enzymes whose rôles range from the utilization of iron to the pigmentation of skin. In many marine invertebrates, copper is an important constituent of the blood protein haemocyanin (e.g., Rose & Bodansky, 1920; Bannister & Wood, 1971). Although primarily involved in biochemical mechanisms, copper is apparently a structural feature of glycerid polychaete jaws (Gibbs & Bryan, 1980) accounting for ≈1.5% of the dry weight of the jaws and representing up to 67% of the total body burden of copper in the organisms. Ulmer (1977) lists several copper-

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containing metalloenzymes (e.g., cytochrome oxidase, lysine oxidase, and tyrosinase). Key biochemical functions also include mitochondrial activity, collagen metabolism, and melanin formation. Signs of deficiency in man include Menke’s syndrome, anemia, and leukopenia. Copper is a member of a group of metals known as transition elements, whose unique property lies in the ability to form strong complexes with ligands or molecular groups (electron donors) of the type found in proteins. Enzymes in which transition metals are tightly incorporated and in which the metal is responsible for the configuration of the protein are called metalloproteins. When the metal is removed the protein loses its capacity to function as an enzyme (Frieden, 1972). The metal can also act in an enzyme substrate system, either as an integral part of the enzyme or as a bridge between the enzyme and the substrate. Copper may also act as a substrate activator by altering the spatial conformation of the substrate. The essential nature of both copper and iron to every form of life is due to their action as prosthetic groups in enzymes and proteins (Frieden, 1974). Frieden points out that the superior chelating properties of the cupric ion are probably the reason that copper proteins developed. Hatanaka & Egami (1977) were able to form certain amino acids from formaldehyde and hydroxylamine in a modified sea medium enriched with transition metal ions including copper. Their results suggest the possibility that natural selection of pre-biotic organic molecules occurred as a result of the nature of environmental catalyzers in the course of chemical evolution. Even when it acts indirectly, copper plays an important rôle. Examples of this include the production of haemes and their incorporation into haemoglobin, and the production of chlorophyll (e.g., O’Kelley, 1974). Copper is, in essence, involved in the processes which lead to the production of all porphyrins (White, Handler & Smith, 1973). It is present in haemocyanin in a nonporphyrin pyrole containing prosthetic group. The prosthetic group provides a stable organometallic complex to which oxygen can be reversibly bound (Redfield, 1934). The chemical properties of haemocyanin are due largely to its protein nature (Redfield, 1934) although the ability to bind oxygen is due to the coppercontaining prosthetic group (Redmond, 1968; Hughes, 1972). Decleir, Vlaeminch, Geladi & Van Grieken (1978) found a high concentration of protein-bound copper and zinc in the branchial gland of the cuttlefish Sepia officianalis and suggested that the gland played an important rôle in trace element metabolism. They also found a low molecular weight copper- and zinc-containing fraction of water soluble material from the gland which they suggested was a small protein or peptide involved in the synthesis of haemocyanin, forming enzymes such as superoxide dismutase (e.g. Shatzman & Kosman, 1978) or the metal-binding agent metallothionein. Salinity has been found to affect oxygen consumption through osmotic stress (Weiland & Mangum, 1975; Gilles, 1977). Gilles found that increasing salinity caused a decrease in oxygen consumption and postulated that this could be a result of osmotic stress producing a decrease in the copper containing proteins in the blood. This is supported by Boone & Schoffeniels (1979) who found that the copper concentration in the haemolymph of shore crabs (Carcinus maenas) adapted to low salinity was twice that found in animals adapted to sea water. They noted, however, that the total body copper content remained constant under both conditions suggesting that copper is stored endogenously for haemocyanin synthesis. Their work also supported the earlier

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suggestion of Schoffeniels (1976) that an increase in the copper content of the blood of the shore crab was a direct reflection of haemocyanin synthesis. Horn & Kerr (1963) found no correlation between size and mean serum protein or size and copper concentration in the blue crab, Callinectes sapidus. Copper has been shown to be essential for the growth and metamorphosis of a wide variety of plants and animals. Noda & Horiguchi (1971) were able to show that in a medium deficient in trace metals the marine alga Porphyra tended to produce carbohydrates in preference to proteins. Thind & Maden (1973) found that copper was one of four trace elements essential for the growth and sporulation of two species of fungi. Manahan & Smith (1973) discussed copper as a growth requirement for algae. Copper has been shown to be important for the successful growth of dinoflagellates (e.g., Anderson & Morel, 1978) and is present in artificial media for determining the biological effect of heavy metals (e.g., AQUIL, Morel, Rueter, Anderson & Guillard, 1979). It has also been shown to be important for the culture of yeasts and bacteria (e.g., Babitskaya, 1978). Copper has been found in the sperm motility initiating factor released from the eggs of the horseshoe crab, Limulus polyphemus, (Clapper & Brown, 1980) and has also been suggested as important in the growth and metamorphosis of animals such as tunicates (Grave, 1941; Bertholf & Mast, 1944; Glaser & Anslow, 1949), oysters (Prytherch, 1931, 1934; Korringa 1941), echinoderms (Lillie, 1921), freshwater prawns (Shakuntola, 1976), and barnacles (Bernard & Lane 1961). The metal is also used as a dietary supplement for fish such as the Red Sea bream, Chrysophrys major (Sakamoto & Yone, 1978) and the prawn, Penaeus japonicus (Deshimaru & Yone, 1978). Fletcher & King (1978b) found that blood plasma concentrations of copper in sockeye salmon (Oncorhynchus nerka) decline during the non-feeding migration to the spawning grounds and suggested that copper may be stored in the body for oocyte development well in advance of spawning, the embryo presumably deriving part or all of its copper from metal stored in the oocyte prior to fertilization. Fletcher, Kiceniuk, King & Payne (1979) also noted a reduction in the blood plasma concentrations of copper in the cunner, a marine fish (Tautogolabrus adspersus), after exposure to an oil slick for six months. They suggested that copper uptake across the gut wall may have been affected by the oil. Much of the work that has been done on growth requirements has been with growth media containing one or more natural or synthetic complexing agents that serve as metal buffers (e.g., Provasoli, 1963). These agents cause the metal to occur in a number of chemical states and the biological availability varies with the chemical state as well as the concentration (e.g., Hutner, Provasoli, Schatz & Haskins, 1950; Hutner, 1972; Lewis & Whitfield, 1974). The concentrations identified as “required” may thus be a result of natural or added metal buffering agents as well as the requirements of the organism. Goldman (1965) suggested that copper is probably never a limiting nutrient in natural conditions because it is available in sufficiently large quantities in most aquatic systems. This is supported by Bender & Gagner (1976) who suggested that copper and nickel are not limiting nutrients in the Sargasso Sea. Boyle & Edmond (1975) found that copper is enriched in the surface waters of the Antarctic upwelling areas. From the correlation with nitrate they suggested, however, that copper might be a limiting nutrient, that levels in the surface water in low and mid-latitudes should be close to zero, and that maximum values in the deep Pacific should not exceed 5 nmol·kg−1.

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DETRIMENTAL EFFECTS The detrimental effects of copper have been discussed in a number of reviews (Hueckvan der Plas, 1972; Steinberg, 1972; Waldickuk, 1974; Karrick & Gruger 1976; McIntyre, 1976; Braunstein, Copenhaver & Pfuderer 1977; Reish et al., 1977; Schmidt, 1978a,b; So, 1978; Eisler, 1979b; Eisler, Rossoll & Gaboury, 1979; Hodson, Borgmann & Shear, 1979). General review papers are available that discuss sources of metals, chemical changes in the environment, and the biological importance of metals. These papers fall into three categories: response of estuarine organisms to changes in water quality (Lockwood, 1978), biological cycles for toxic elements in the environment (Wood, 1974), and the behaviour of trace metals (Ruivo, 1972; Wood & Goldberg, 1977; Spear & Pierce, 1979; Zingaro, 1979). Frazier (1972) lists the research required to obtain an adequate knowledge of heavy metals in Chesapeake Bay, a list that could be universally applied. Concern about the effect of pollutants, including copper, has been expressed by government, industry, and academia (e.g., Raymont & Shields 1963; Parker, 1974; Alabaster, 1976; Davey, 1976a; Hura, Evans & Wood, 1976; Castilla & Nealler, 1978). This is well shown by the National Oceanic and Atmospheric Administration reports to the U.S. Congress on ocean pollution, overfishing, and offshore development (N.O.A.A., 1975, 1977, 1978); the symposium held jointly by the U.S. Environmental Protection Agency and the Soviet Academy of Sciences (Turekian & Simonov, 1978); The American Chemical Society Symposium (e.g., Hoigne & Bader 1978) which included papers on recent developments in the environmental chemistry of some trace elements; and finally, the series of International Symposia on Environmental Biogeochemistry (e.g., Loring, 1978c). The two volume series edited by Nriagu (1979c) and entitled Copper in the Environment provides a series of papers dealing with the chemical and biological properties and effects of copper in terrestrial, freshwater, and marine environments. Bryan (1976) provides an excellent review of heavy metal contamination in the sea and discusses some of the complexities of the interactions between organisms and metals such as copper. The 1977 report by the U.S. National Oceanic and Atmospheric Administration to the U.S. Congress on ocean pollution, overfishing, and offshore development focuses on six areas of interest, amongst which are the heavy metal problems associated with sludge dumping in the New York Bight, the relationships of heavy metals and selected organisms, and some of the environmental questions raised by deep-ocean mining. Toxic effects of metals are many (e.g. Vernberg & Vernberg, 1974). Ulmer (1977) suggests that metals may serve as enzyme inhibitors, that they may inhibit oxidative phosphorylation, alter membrane permeability, impair protein synthesis, or distort nuclei acid structure. Alayse-Danet, Charlou, Jezequel & Samain (1979) have found that high levels of copper (2 mg·l−1) affect both amylase and trypsin activity. Werringloer, Kawano, Chacos & Estabrook (1979), as well as others, have found that copper, either as cuprous sulphate or chelated, interacts with the microsomal transport system and modifies some of the electron transport reactions. Copper has been shown to have deleterious effects on cell division (Kanazawa & Kanazawa, 1969), photosynthetic rate (McBrien & Hassall, 1967; Kamp-Neilsen, 1969;

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Steemann Neilsen, Kamp-Neilsen & Wium-Anderson, 1969; Gross, Pugno & Dugger, 1970; Steemann Neilsen & Wium-Anderson, 1971; Zingmark, 1972; Patin & Tkachenko, 1974), and growth (Den Dorren De Jong, 1965; McBrien & Hassall, 1967; KampNielsen, 1969; Mandelli, 1969; Arnold, 1971; Steemann Neilsen & Wium-Anderson, 1971). Other studies have dealt with the ability of copper to reduce “blooms” of algae (Muehlberger, 1969; Telitchenko, Tsytsarin & Shirokova, 1970; Sawyer, 1971) and zooplankton populations (e.g., McIntosh & Kevern, 1974). Steemann Neilsen & WiumAnderson (1972) found an overall depression of photosynthesis by 3 and 6 µg·l−1 copper but with most of the effect being apparent during the afternoon. Clendenning & North (1960) and North (1964) found a reduction in photosynthesis in the giant kelp Macrocystis pyrifera with 0.1 mg·1−1 copper. North (1964) also notes that with 1 mg copper·1−1 the blades exuded a yellow substance and that the kelp had a “new mown hay” odour. This work was done to determine the effects of discharged wastes from a hyperion outfall and North suggests that, at the time of writing, at least in the San Diego area, introduced heavy metals were unimportant to the kelp. Rajendran, SumitraVijayaraghavan & Wafar (1978) examined the effects of copper and some other metal ions on the photosynthesis of “microplankton and nannoplankton”, using the dark and light bottle technique with 14C. Although they do not mention the form of copper that was used they do indicate that a small addition of copper (8 µg·l−1 for microplankton and 4 µg·l−1 for nannoplankton) enhanced productivity although photosynthesis dropped abruptly with increasing copper. They also note that inhibition due to copper was much more than with zinc, molybdenum, iron, or cobalt. Reduction in photosynthesis is not only attributed to copper however, Davies & Sleep (1979) suggest that zinc pollution from river drainage may be responsible for a reduction in photosynthesis in some British coastal waters although they note that natural chelating agents may play a rôle in reducing toxicity of zinc in much the same manner as they do the toxicity of copper in phytoplankton cultures. Photosynthesis is inhibited by abnormal concentrations of copper which act on the photosynthetic pigments (McKee & Wolf, 1963; Kamp-Nielsen, 1969; Steemann Neilsen et al., 1969; Steemann Neilsen & Wium-Anderson 1971; Niemi, 1972) although the mechanism of action is not known. The inhibition of photosynthesis as a result of the activity of copper causes an inhibition of growth (Mandelli, 1969). Lehman & Vasconcelos (1979) found that copper concentrations of 1000 µg-at.·l−1 were necessary to reduce photosynthesis by 50% in Cylindrotheca clostridium despite a 50% growth inhibition at 10 µg-at.·l−1, with respiratory rates paralleling photosynthetic rates with increased copper stress. They also noted changes in lipid distribution with respect to both polar and non-polar lipids. As reported by Lehman (1979), sterol esters, sterols, galactolipids, phosphatidylethanolamine, sterol glycosides, phosphatidic acid, and free fatty acids were reduced in copper stressed cells. The effect of copper on lipids appears to be widespread, Mizushima, Takama & Zama (1977) found that lipid deterioration in fish flesh homogenates was increased by copper as well as iron. Harrison, Eppley & Renger (1977) found several acute effects of copper with phytoplankton cells. These effects included inhibition of nitrate uptake, reduction in photosynthetic carbon assimilation, reduction of synthesis of nitrate reductase, as well as cell disruption and loss of accumulated ammonium. Riisgård (1979) and Riisgård,

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Neilsen & Søgaard-Jensen (1980) found that copper concentrations of ≥2 mg·l−1 inhibited regulation of cell volume in the marine flagellata Dunaliella marina. Thomas, HolmHansen & Seibert (1977) and Thomas et al. (1977b) found that phytoplankton in the presence of elevated copper levels excreted high amounts of 14C-labelled organic matter while Georing, Boisseau & Hattori (1977) found that silicic acid uptake in phytoplankton was inhibited by copper additions while Bankston et al. (1979) found that silicon enrichment of culture medium reduced copper uptake in several groups of phytoplankton. Morel, Rueter & Morel (1978) noted that the toxicity of copper to the diatom Skeletonema costatum was a function of the silicic acid concentration of the medium. Harrison & Davies (1977) examined the effect of copper on the flux of nitrogen through a planktonic marine food chain and noted that there was a chronic inhibition of zooplankton feeding which, along with the reduction in nitrate uptake, produced a change in the community that could be attributed to the effects of copper. Bentley-Mowat & Reid (1977) noted a species difference in the response of batch cultured phytoplankton to the addition of 10−4M copper, cadmium, or lead. Growth of several species did not appear to be affected while one, Ditylum brightwelli, underwent osmotic disturbances with both 10−5 and 10−4M copper, with a swelling of the cell. Membrane permeability appears to be affected by copper in a wide variety of organisms. Khovrychev, Ivanova & Taptykova (1974) for example, found this in the yeast Candida utilis and noted that the content of protein and RNA decreased in the cells while the total content of lipids increased, possibly as a result of a change in the permeability of the cell membrane. With massive amounts of copper (50–1000 µg·l−1) Eyles (1975) showed that photosynthesis, cell division, growth rates, and respiration were affected in a diatom (Cylindrotheca closterium var. californica) and a dinoflagellate (Amphidinium carterae). Pace, Ferrara & Del Carratore (1977) found that 5 mg Cu2+·1−1 was lethal to Dunaliella salina while, at lower concentrations, the copper affected cell numbers as well as pigment composition. Berland, Bonin, Guerin-Ancey & Arlhac (1977), working with the diatom Skeletonema costatum, found that the division rate was sensitive to increasing copper concentrations, that the C: N ratio remained unchanged at sublethal concentrations, and that cell volume was only slightly affected. Saifullah (1978), working with several marine dinoflagellates, found that copper addition inhibited growth but that the effect was more pronounced in semi-continuous culture than in batch culture. Kanazawa & Kanazawa (1969) demonstrated that copper interacts with membrane ATPase to inhibit cell division in Chlorella. Gächter (1976) studied the effects of several heavy metals on phytoplankton and found that photosynthesis was not adversely affected up to a copper concentration of 5×10−9M, but that there was a synergistic effect from other metals. Gächter also noted seasonal variation in the toxic effects of heavy metals and attributed it to changes in phytoplankton composition. Overnell (1976) measured the effect of heavy metals on “the maximum rate of light-induced evolution of oxygen” in cultures of several unicellular marine algae (At they a decora, Brachiomonas submarina, Dunaliella tertiolecta, Isochrysis galbana, Monochrysis lutheri, Phaeodactylum tricornutum, and Skeletonema costatum). S.costatum and Attheya decora were found to be sensitive to both copper and mercury ions. Phaeodactylum tricornutum, in contrast, was insensitive to copper after 15 min of dark pre-incubation. Overnell also noted that calcium, magnesium, and potassium ions

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afforded protection for P.tricornutum against copper in half-strength sea water. Ibragim & Patin (1976) examined the effect of mercury, lead, cadmium, and copper on primary production and phytoplankton in some coastal regions of the Mediterranean and the Red Sea. They found (p. 589) that “…significant changes can take place in the rate of photosynthesis and in the species composition of phytoplankton communities under the influence of concentrations of metal close to their natural levels in seawater”. They also noted that, in the range of concentrations studied (1, 10, 100, 1000 µg·1−1), mercury was the most inhibiting followed by copper, cadmium, and lead. Overnell (1975), however, found that copper did not affect photosynthesis in the freshwater alga Chlamydomonas reinhardii as much as cadmium, methyl mercury, and lead. Hopkin & Kain (1978) examined the effects of some pollutants on the survival, growth, and respiration of the alga Laminaria hyperborea under culture conditions. Survival of the germinating gametophytes was reduced by 0.1 mg copper·dm−3, the production of sporophytes from the garnetophyte generation was delayed by 0·01 mg copper·dm−3, and the respiration of frond discs was reduced by 25 mg copper·dm−3 although there was a slight stimulatory effect at a copper level of 10 mg·dm−3. Christensen, Scherfig & Dixon (1979) examined the effects of manganese, copper, and lead singly and in combination, on two algal species (Selenastrum capricornutum and Chlorella stigmatophora). With artificial media they found a reduction in total cell volume with high levels of copper (85 µg·1−1 for Selenastrum, 70 µg·1−1 for Chlorella) added singly. There was a synergistic reaction between manganese and copper and an antagonistic reaction between copper and lead. Gauthier, Bernard & Aubert (1976) examined the effect of paired metals, including copper, on production of a bactericidal antibiotic produced by a marine species of Alteromonas and found that the action varied with different metals. The synergistic activities were found to be controlled by salinity and organic matter. Khovrychev, Ivanova & Taptykova (1974) noted that “respiration” of the cells of the yeast Candida utilis was inhibited when copper was introduced in the exponential phase of growth. Copper-containing herbicides (e.g., cutrine) are frequently used for control of plant growth in freshwater ponds (e.g., Hestand, Carter & Royals, 1977). One of the results of this is to affect the species composition and the succession of species in the pond after herbicide treatment. Myers & Cooke (1978) attributed the succession to nutrients released from decaying vegetation by the herbicide treatment. Meyer (1978), however, found that copper and zinc were two metals that could not be correlated with any obvious algal succession in a study of the effects of heavy metals on algal populations in a freshwater reservoir. With mixed cultures of marine algae, Fielding & Russell (1976) found that the species responded more positively to copper and at lower concentrations than they did in unialgal cultures. They also found that species interaction changed the species response. Ulothrix, for example, enhanced the growth of Ectocarpus at 100 µg Cu·dm−3. Ryther & Sanders (1979) and Sanders, Ryther & Batchelder (1981) examined the effects of environmental stresses on the species composition of phytoplankton populations, using free chlorine, copper, and increased temperature as stressing agents. Copper-exposed cultures exhibited a decreased total population size and a lower diversity, with a similarity index significantly different from that of the control. Thomas et al. (1977a,b, 1980), working with the CEPEX bags and laboratory cultures, found that species

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composition changed under copper stress, that centric diatoms were most sensitive, and were reduced while pennate diatoms and microflagellates were more resistant. The nature of the copper species is suspected to be important in controlling species succession in upwelled waters (e.g., Lewis, 1977); the cupric ion concentration playing an important rôle in survival of phytoplankton species (e.g., Anderson & Morel, 1978). Large populations of unicellular algae have been noted in all the world’s oceans. These “blooms” are most frequently observed in inshore areas and have been associated with a number of events. Hirayama & Iizuka (1975) for example, found that anaerobically decomposed products of bottom mud were stimulating to the growth of Gymnodinium and suggest that this could be due to the supply of inorganic nutrients and some “unidentified stimulants”. Prakash (1975) as well as others, associate “red tide” outbreaks with land drainage. Prakash points out that dinoflagellate blooms occur at times of high organic loads in sea water and that pollution of inshore coastal waters, especially by sewage, may increase the frequency of the blooms. The presence of both iron and humic compounds, which act as complexing agents, have been suggested to be key components of the land drainage (e.g., Martin, Doig & Pierce, 1971). Pingree, Holligan & Head (1977) discuss the survival of dinoflagellate blooms in the western English Channel in terms of hydrographic conditions. They suggest that injection of nutrients occurs in particular zones of tidal mixing (Pingree, Pugh, Holligan & Forster, 1975). There is thus a mechanism for effectively concentrating biologically important factors (“humic compounds”, iron, copper) and important nutrients (e.g., nitrates) and for providing an environment suitable for the population increase necessary for a plankton bloom. Anderson & Morel (1978) show copper toxicity to Gonyaulax tamarensis, a bloomforming dinoflagellate, to be a function of cupric ion concentration. They suggest that, with natural levels of copper, organic complexation would be necessary before the species could successfully compete with other algal species in coastal waters. Later work (Anderson & Morel, 1979) gives further details of the importance of these agents. The conditions produced by the hydrographic “mechanism” found by Pingree et al. (1975), combined with the run-off factors mentioned by Prakash (1975) and the decomposition of organic matter found by Hirayama & Iizuka (1975), may provide localized regions of complexing agents adequate to provide the level of cupric ion required by bloom-forming dinoflagellates such as G.tamarensis. Copper has been shown to produce a number of behavioural, histological, and physiological anomalies in both plants and animals. Histological effects of copper toxicity have recently been noted in the freshwater plant Lyngbya nigra where concentrations above 0.8 µM caused formation of large numbers of separation discs, the trichomes contracted longitudinally, and the cells became swollen and constricted at the cross walls as well as turning yellowish and losing photosynthetic pigments (Gupta & Arora, 1978). Mueller (1979) found that what were termed “low” concentrations of cadmium, copper, and lead produced a change in the behaviour of several intertidal organisms in the outer estuary of the Elbe River. The rhythmic movement of valves of the freshwater mussel Anodonta cygnea was affected by 10−8 g CuSO4.1−1 (Salanki & Varanka, 1976) while Manley & Davenport (1979) determined the level of total copper at which this occurred in a series of marine bivalves.

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Modiolus modiolus varied from the normal pattern at a concentration of 0.141 mg·1−1 (±0·108) added copper, Mytilus edulis at a concentration of 0.020 mg·1−1(±0·01), Crassostrea gigas at 0.085 mg·1−1(±0·025), Anadara senilis at 0.120 mg·1−1(±0.012), and Modiolus demissus at . McGreer (1979a) noted that the burrowing behaviour of Macoma balthica was inhibited in heavy metal contaminated sediments near a sewage outfall and that the relationship between burrowing behaviour and heavy metal concentration was significant for mercury and cadmium but not copper. He also noted, from laboratory work, that the clams exhibited an active avoidance response to sediment containing the highest levels of metals. The burrowing response of the marine bivalve Tellina tenuis was changed by concentrations of copper >250 µg·1−1 (Dow, Bell & Harriman 1975). Feeding rates of two calanoid copepods (Pseudocalanus sp. and Calanus sp.), a euphausiid (Euphausia sp.) and a ctenophore (Pleurobrachia sp.) were reduced by added copper (Reeve, Gamble & Walter, 1977b). Martin, Piltz & Reish (1975) noted that the addition of copper caused a decrease in the ability of the marine mussel Mytilus edulis to produce byssal threads. Inamori & Kurihara (1979a) examined the effects of copper on the brackish-water polychaete Neanthus japonica and found that “…in the case of cupric ions survival was satisfactory at 1 mg/l but fell to zero at 3 mg/l, while the weight gain and food conversion efficiency declined at 0.5 mg/l”. Although it is doubtful that all of the copper was in the cupric ion state after addition to the water, especially with survival at the 1 mg·1−1 level, it is of interest to note that the sublethal effect of copper on the polychaete was associated with feeding, even although copper was added as CuCl2. Reish (1978) found that the sensitivity of polychaetous annelids to copper varied with the species while Siebers & Ehlers (1979) noted that 0.25 mg·1−1 copper significantly reduced 14C-glycine uptake across the body surface in the oligochaete Enchytraeus albidus. Uptake of amino acids by the gills, digestive gland, and mantle of the mussel, Mytilus galloprovincialis, was reduced to 5–10% of the control value after a week of exposure to 0.08·1−1 (Viarengo et al., 1980a,b). Jørgensen & Jensen (1977) commented that “… the effects of copper (II) chloride on Artemia hatching rate are seen at concentrations of the same level as found in natural sea water”. They concluded that even a minor discharge of copper into the nearshore environment gives reason for concern. MacInnes & Calabrese (1978) found that the embryos of the oyster, Crassostrea virginica, were more susceptible to metal toxicity at some temperatures than others, toxicity was higher at 20 and 30°C than at 25° C. Keeney (1974) found that in C.virginica NADH oxidation by MDH was inhibited by several metals, in the order . Baker (1969) looked for histological changes in the winter flounder, Pseudopleuronectes americanus. High and medium levels of copper resulted in fatty metamorphosis of the liver, kidney necrosis, obstruction of the haematopoetic tissue, and gross changes in the gill architecture. Low levels of copper caused a reduction in the glandular structures associated with the gill lamellae and epithelial tissue. Copper was shown to affect the central nervous system and kidney functions, act as a mucous coagulant, and induce secretory cells to become chlorotic. Haemopathological changes were found in the mullet, Liza macrolepis, by Helmy, Lemke, Jacob & Oostdam (1978) after 96-h exposures to copper. Changes in leucocytic profile and the percentage eosinophils appeared to be associated with pathological changes caused by increasing

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copper and mercury suggesting that blood measurements could be used to diagnose high levels of these metals. Schreck & Lorz (1978) noted a marked, dose-dependent association between copper and the level of serum cortisol in the coho salmon, Oncorhynchus kisutch. Salmon exposed to elevated levels of copper also had depressed serum chloride levels and reduced survival when challenged with salt water. Sherwood (1977) compared levels of seven trace metals, including copper, in the liver tissue of the Dover sole, Microstomus pacificus, and starry flounder, Platichthys stellatus, with the frequency of fin erosion and found a variation between two regions suggesting that the association between copper and fin erosion, if any, is affected by local conditions. Eisler & Gardner (1973) found that the renal and lateral line canal of the teleost fish Fundulus heteroclitus (mummichog) developed lesions when the fish was subjected to copper concentrations of 1 mg·1−1 and that the epithelia lining the oral cavity was necrotized by the caustic action of high levels of zinc (60 mg·1−1) and copper (8 mg·1−1). They also noted a synergistic effect of copper and zinc, related to the individual toxic properties of metals at different anatomical locations or sites of activity. Delhaye & Cornet (1975) noted that elevated copper levels caused an increase in mortality of Mytilus edulis during reproduction and attributed it to an increase of the animal’s metabolism. Oxygen consumption rates of the mud snail, Nassarius obsoletus, were found to be depressed by copper, arsenic, silver, and zinc individually and copper and cadmium together (MacInnes & Thurberg, 1973). Although cadmium caused an increase in oxygen uptake, combination with copper produced a decrease greater than that by copper alone. Kerkut & Munday (1962) found that certain copper salts increased the respiration rate of isolated tissues of the crab, Carcinus maenus. According to these authors, “coppersodium-potassium tartrate” at all concentrations initially produced an increase in most tissues but then a decrease. At a concentration of 100 meq all tissues respired at a rate <50% of normal. The oxygen uptake of cyprid larvae of Balanus amphitrite niveus was increased by

low concentrations of copper and inhibited by high concentrations (Bernard & Lane, 1963). Similar results were found for the shrimp, Caridina rajadhari, (Chinnayya, 1971). Mueller (1979) measured oxygen uptake in three organisms (the shrimp, Crangon crangon, the polychaete, Nereis diversicolor, and the goby, Pomatoschistus microps) “…typical of the biocoenosis of the intertidal flats at the mouth of the Elbe-River”. Oxygen consumption was affected by Cu2+ (as CuSO4) levels as low as 5 µg·1−1. In both a publication (Cardeilhac et al., 1979) and a paper abstract (Cardeilhac, Yosha & Simpson, 1979) the effect of copper poisoning on the marine teleost fish, Archosargus probatocephalus, was reported to produce a potassium intoxication caused by cell damage and failure of osmoregulation by both gills and kidneys. Copper has been found to affect osmoregulation in several crustaceans. Thurberg, Dawson & Collier (1973) found that crabs (Carcinus maenas, Cancer irroratus) exposed to copper exhibited a loss of osmoregulatory function with increasing copper

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concentration until the normally hyperosmotic serum became isosmotic with the surrounding medium. In contrast, they found that CuCl2 had no effect on gill tissue oxygen consumption. (There was a suggestion that the species of copper used may have been important as other copper salts were highly inhibitory.) Price & Uglow (1979) examined the effects of copper and other metals on development and mortality within the moulting cycle of the shrimp, Crangon crangon, noting that the normal uptake of water immediately after ecdysis allows the introduction of trace metals at a physiologically poor time. Because of the increased permeability of the cuticle at immediate post-moult, more work is required to maintain haemolymph isotonicity and the presence of appreciable quantities of copper and zinc ions would add to this workload as well as decrease the efficiency of the regulating tissue. Osmoregulation is also affected by copper in a number of fish: coho salmon (Lorz & McPherson, 1976), the channel catfish and golden shiner (Lewis & Lewis, 1971), and the striped bass (Courtois, 1976). Exposure to elevated levels of copper causes breakdown of the cell membrane. Gardner & LaRoche (1973) found lesions in the olfactory organs and lateral line canals in the head of the adult niummichog, Fundulus heteroclitus, and Atlantic silverside, Menidia menidia, while Young (1978) noted sloughing epithelia, vacuolation, hyperplasia, and necrosis in transport epithelia of certain polychaete worms. Another result of heavy metal exposure in some organisms is the storage of excess metals, primarily copper and zinc, in cell structures. Ruddell & Rains (1975) discuss the relationship between zinc, copper, and the basophils of two oysters (Crassostrea gigas and C.virginica), noting that the basophils act as storage chambers for the metals (see also George et al., 1978). Walker (1977) as well as others have found copper granules in parenchyma cells of the prosoma of barnacles. Ruddell & Rains (1975) as well as Walker (1977), found that the metal was associated with organic material. In the case of the oysters, the organic material appears to be phenolic in nature (Ruddell & Rains, 1975) while in the case of the barnacle the nature of the organic matter is unknown. In both the oysters and the barnacle, the organisms are removing copper (and zinc) only as required; Ruddell & Rains point out that the metal requirements of the oysters will be governed by the turnover rates of the zinc and copper metalloenzymes and the basophils. They state that “…until the mechanisms governing the turnover of oyster basophils are fully elucidated, one should not ascribe large amounts of zinc and copper in oysters to the effects of mining or industrial pollution”. The eel, Anguilla anguilla, is found in both fresh and salt water during its life history. Rodsaether et al. (1977) have found that when eels are in fresh water they may establish a commensal association with the bacterium Vibrio anguillarum. When eels were exposed to copper-contaminated fresh water (30–60 µg Cu·1−1) they died with signs of vibriosis, an increase in the concentration of Vibrio anguillarum. These authors suggested that the effect of the copper was to turn the commensal relationship into one of pathogenicity. Thomsen (1980), however, noted that copper did not seem to be an initiating factor in developing vibrio infection in Anguilla anguilla. Stevens (1977) found that the immune response of coho salmon to Vibrio anguillarum was significantly reduced when the fish were exposed to 18.2 µg copper·1−1 and Hetrick, Knittel & Fryer (1979) found an increased susceptibility of rainbow trout (Salmo gairdneri) to infectious haematopoetic necrosis virus after exposure to copper. They noted that although the level of copper

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influenced mortality rates the length of the exposure did not. The mechanism responsible for the rainbow trout susceptibility to the virus was not determined although the occurrence of a marked, copper-produced, dose-dependent elevation in several hormones, including corticoids and serum cortisol, suggested possible hormonal effects. The biological effect of copper is different during different parts of the life history of marine organisms. Generally speaking, the early life history stages are more sensitive than are the later stages. Harrison (1977a) found that the average growth of copper-spiked (50 µg·1−1) cultures of the giant kelp Macrocystis pyrifera were inhibited 10% after one week, 30% after two weeks, and 40% after four weeks. Hopkin & Kain (1978) found that the zoospores and gametophyte stages of Laminaria hyperborea were more tolerant than the sporophyte stage. Stromgren (1980) found significant reductions in growth of four species of intertidal fucoid algae (Pelvetia canaliculata, Fucus spiralis, F.vesiculosus, and F.serratus) with levels of copper ranging from 12–50 µg·1−1. Weiss (1948) found that barnacles that settled and grew on copper antifouling paints, copper alloys, and zinc grew in an abnormal fashion with the base deeply scalloped, the plates of the shell distorted and with excessive calcification of the base although they remained loosely attached to the substratum. Harrison (1977a) noted increased mortality in the early life history stages of Crassostrea gigas at copper concentrations as low as 5 µg·1−1 and commented that the early life history stages of the oyster are extremely sensitive to increased concentrations of copper. Harrison, Emerson & Rice (1977) also noted that in addition to increased mortality, elevated levels of copper caused an increase in the number of abnormal larvae, a decrease in size at hatching as well as the percentage hatching, changes in behaviour, changes in oxygen consumption, and changes in heart rate. MacInnes & Calabrese (1979) noted that low levels of copper may increase the stress on the recruitment of oyster embryos during periods when environmental conditions of temperature or salinity are extreme. Blaxter (1977) reported that newlyhatched larvae of plaice (Pleuronectes platessa) and herring (Clupea harengus) exhibited a high mortality at 1000 µg copper·1−1 (an unreasonably high level). He also noted that feeding by yolk-sac plaice was totally inhibited by 90 µg copper·1−1, a level more than an order of magnitude higher than what is found in inshore waters. Rice & Harrison (1978a) found that the embryonic stage of the northern anchovy, Engraulis mordax, was the stage most sensitive to copper, with a 12-h LC-50 of 200 µg copper·1−1. Voronina & Gorkin (1978) examined the effect of copper on the early life history of Tilapia, an important aquaculture fish. They found a high mortality of embryos with copper concentrations between 1.5–2.5 mg·1−1, a retarding of hatching at concentrations between 0.3–1.0·mg 1−1 and hypertrophy of the organs associated with blood circulation at levels of 0.01–0.1 mg·1−1. Ozoh (1980) incubated the eggs of the zebrafish, Brachydanio rerio, with and without the shell membrane, in copper- and lead-enriched media and found that the shell membrane appeared to provide limited protection against the toxicity of copper. Reish, Piltz & Martin (1974), working with the polychaete, Capitella capitata, found that copper levels of 0.01–0.05 mg·1−1 caused the production of a bifurcate larvae. Although the incidence was low (0.6–0.9%), they noted that only one generation was required to induce the abnormal larvae in the presence of sublethal amounts of copper while two generations were required to induce the abnormal larvae in the presence of sublethal amounts of zinc and detergents.

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Some terrestrial and aquatic organisms exhibit the ability to grow in high levels of copper. Hogan, Courtin & Rauser (1977a, b) found that two clones of the grass Agrostis gigantea from a mine waste site would not grow in soils in which copper was not readily available. Brown & House (1978) found a copper-tolerant clone of Solenostoma crenulatum, growing on mine soil tips. Gartside & McNeilly (1974) suggested that tolerance to copper in plants is evolved and heritable, but that there is considerable interspecific variation. Thomas et al. (1977a) found that copper and mercury decreased algal productivity but that recovery occurred with those species which were resistant to the metals. A similar result was obtained with brackish- and freshwater bacteria (Singleton & Guthrie, 1977). Adaptation to elevated levels of metals has been reported in a number of organisms. Stebbing & Hiby (1979) used a copper as a growth inhibitor of the colonial hydroid, Campanularia flexuosa, and found “…that the inhibitory effect of copper is effectively antagonised between 1 and 10 µg/l, and since the organism responds similarly to other inhibitory agents, it appears that the counteractive capacity of the growth control mechanism may be a significant factor influencing the effects of toxicants on growth”. Nereis diversicolor collected from sediments with high metal loadings can tolerate increases in metals greater than those from sediments with low metal loadings (Anonymous, 1972; Bryan, 1974). Moraitou-Apostolopoulou (1978) and MoraitouApostolopoulou & Verriopoulos (1979) found that Acartia clausi from a polluted region had a higher tolerance to copper and a higher fecundity at sublethal levels than did a population of the same species from a relatively unpolluted region. The same trend has been reported in algae (Foster, 1977) and bacteria (Timoney, Port, Giles & Spanier, 1978). Jensen & Rystad (1974) found that a clone of the chain-forming diatom, Skeletonema costatum, isolated from a zinc-polluted environment tolerated elevated levels of zinc better than a clone isolated from an environment with low levels of zinc. The tolerance seems to be metal specific as the zinc-tolerant clone was not more tolerant to copper than the zinc-intolerant clone. In contrast, Stokes, Hutchinson & Krauter (1973) noted that clones of freshwater algae belonging to the genera Chlorella and Scenedesmus, from metal-rich lakes, were not only tolerant to high levels of nickel and copper, but also to silver even although silver was not a pollutant in the lakes. The suggestion from these studies is that tolerance can be achieved but that the nature of the organism may dictate the degree of specificity. The idea that a given species may achieve a tolerance of high levels of metals suggests either a selection for the more tolerant members or a change in one or more biochemicalphysiological mechanisms as a result of exposure, or both. The suggestion that populations within species (demes) occur in the marine environment is not new although recent work on the loss of heterozygosity in clonal cultures of Skeletonema costatum and Thalassiosira pseudonana (Murphy, 1978) suggests that these changes can occur in a relatively short time. In combination with the results of a study by Murphy, Guillard, Lee & Brand (1978), on the distribution of electromorphs and growth rate characteristics in isolates of T.pseudonana from the neritic oceanic boundary, it appears not only that changes can occur but also that they are in response to environmental conditions. Murphy & Belastock (1980) found that clones from heavily polluted inshore waters were much less sensitive to pollutants than clones from relatively unpolluted inshore waters. The

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study of Saliba & Krzyz (1976) indicated that the chemical state of the metal may be important in acclimatization. They found that the brine shrimp, Artemia salina, acclimated to copper in the chloride, carbonate, and sulphate form, but not in the acetate form. Increased tolerance of heavy metals can be achieved by the localization of excess metals in a chemical form or state that is not toxic. This may be a mechanism normally used for storage of essential metals. Djangmah (1970) for example, found that Crangon vulgaris stores copper from plasma haemocyanin in the hepatopancraeas during moulting. Djangmah & Grove (1970) noted that the copper content of the blood of C.vulgaris maintained in copper-enriched sea water did not increase whereas the hepatopancreas accumulated copper. Thus, not only is the hepatopancreas the site of copper storage during the moult but is capable of storing excess copper to allow maintenance of normal serum copper levels. Johnston & Barber (1969) found that the heptopancreas of Panulirus interruptus contains 75% of the copper which could be used to produce haemocyanin. The presence of copper granules in Balanus balanoides (Walker, 1977) and the metal-containing basophils of two oysters (Ruddell & Rains 1975) may also be a natural mechanism put to use under conditions of elevated metal concentrations. Coombs & George (1977) found that the vesicles in amoebocytes of “Green Sick” specimens of Ostrea edulis were filled with copper associated with sulphur. They also examined Mytilus edulis that had been exposed to high metal levels and found iron, lead, and zinc in membrane-lined vesicles in the cells of specific tissues. The occurrence of higher levels of metals in organisms at certain times of the year has been suggested (e.g., Shiber, 1979) although the work of Van den Broek (1979) with fish and shrimp and Walker & Foster (1979) with Balanus balanoides suggests that a seasonal change in metal levels may occur, but it may not be readily apparent. Walker & Foster (1979) also found that, with Balanus balanoides maintained in the laboratory in running sea water for nearly 18 months, there was a loss of over 75% of the copper in the body. This infers that, at least under certain conditions, the total burden of metals may be reduced. Part of this could be due to the nature of ingested material. Flegal & Martin (1977), for example, found that inorganic residue from ingested sediment in two rocky intertidal gastropods (Tegula funebralis and Acmaea scabra) and two planktonic copepods (Acartia tonsa and A.clausi) often made up a significant proportion of the elemental concentration of the organism although less so with copper than with other elements. Turpaeva & Simkina (1962) examined the ability of a series of marine organisms to tolerate high levels of copper and found that animals possessing differentiated excretory organs (polychaetes, crustaceans) restored vital functions after excretion of the metal. In animals without differentiated excretory organs (e.g., bryozoans, hydrozoan coelenterates) restoration was through the regenerative capacity of that part of the organisms not being poisoned (hydrozoans) or by resting bodies (bryozoans). In the case of bryozoans, the larval form was found to be able to excrete copper because of the presence of a protonephridial type excretory structure while the adult could not excrete copper because of the loss of the excretory structure during metamorphosis. Although excretory organs or new growth may be involved in the tolerance of aquatic organisms to copper it does not explain the mechanism of the involvement. Silver,

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Schottel & Weiss (1975) found that bacterial resistance to toxic metals was determined by extrachromosomal R factors. Jones & Roche (1966) found what they believed was a mucopolysaccharide-copper complex produced by a copper-tolerant bacterium and that copper and other metals were concentrated in the capsular material outside the cell. Olafson & Thompson (1974) found that cadmium-binding proteins could be isolated from liver homogenates of the Atlantic grey seal, Pacific fur seal, and copper rockfish. Although the structure of the isolate was not determined, the properties suggested that it was similar to metallothionein which is a known metal-complexing agent found in several vertebrates (see e.g., Cherian & Goyer, 1978; Parsons & Brown, 1978). Care should be taken in attributing all complexing agents produced in response to metal stress, to metallothionein-like agents because of the wide distribution of agents with similar activity (e.g., Jones & Roche, 1966; Van den Berg, Wong & Chau, 1979). Care should also be taken in interpretation of results. Zevenhuizen, Dolfing, Eshuis & ScholtenKoerselman (1979), in a study of the inhibitory effects of copper on bacteria related to the free ion concentration, found that continuous growth with gradually increasing cupric ion concentration in a static culture lowered the proportion of viable cells. This, with the uptake of copper by all cells, ultimately lowered the cupric ion concentration to tolerable values for the survivor cells. Howard & Nickless (1978) examined heavy metal complexation in periwinkles (Littorina littorea), cockles (Cardium edule), and scallops (Chlamys opercularis) from polluted areas and found three distinct zinc complexes and one copper and cadmium complex from extracts of the periwinkle, with molecular weights reported to be <3000. They considered that the elevated levels of copper in the periwinkle were due to the presence of this compound(s). In addition to the nature of the organism, the environment also exerts some control over the biological effect of copper. Although a discussion of some effects of environmental factors is given elsewhere, it is appropriate to briefly discuss organismenvironment interactions in this section. The fact that the concentration of a metal in the soil does not indicate its biological availability has been known for some time. It has caused the development of tests to identify soils with insufficient biologically available metals for maximum yields to crops (e.g., Lindsay & Norvell, 1978). It has also caused examination of methods of controlling the availability of metals in soils (e.g., Dolar & Keeney, 1974). This study, for example, showed that there was interaction between organic matter, clay, available phosphorus, and metals producing different plant yields in various soil types. The interaction of copper and other metals with organics in both water and sediments has been suggested as important in controlling the biological effect of the metal (Button & Dunker, 1971; Milanovich, Wilson & Yeh, 1975; Stephenson & Taylor, 1975; Chu, 1976; Whitfield & Lewis, 1976; Cave, 1977; Lewis, 1977; UN EIFAC Working Group 1977; Gächter, Davis & Mares, 1978; Jackson & Morgan, 1978; Knezovich & Harrison, 1978; Sunda & Lewis, 1978). The importance of this interaction is suggested by a statement in the report of the UN EIFAC Working Group on the effect of zinc and copper pollution on the salmonid fisheries in a river and lake system in central Norway, “… control of copper pollution based on total copper analysis might be unnecessarily stringent if a major proportion of the copper was present as non-toxic soluble cupro-

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organic complexes”. Sprague (1968), for example, found that the trisodium salt of nitrilotriacetic acid and the disodium salt of ethylenediaminetetraacetic acid reduced the toxicity of copper and zinc to brook trout (Salvelinus fontenalis) and advocated adding some when metal levels became excessive. Anderson & Morel (1979) found that the spring and fall blooms of Gonyaulax tamarensis were caused by a temperature-induced germination of hyponocysts but that the viability of the germling cells was affected by local factors that appeared to control the trace metal chemistry of the environment. The suggestion was that if a natural complexing agent reduces the level of toxic metal species such as the cupric ion, the viability of the germling cells will increase. Young, Gurtisen, Apts & Crecelius (1979a) examined the relationship between the copper-complexing capacity of sea water, as measured by anodic stripping voltammetry (ASV), and the toxicity of copper to the zoeal larval stages of the coon-stripe shrimp, Pandalus danae. They were able to relate the mortality of the larval states to labile copper and the copper-complexing capacity of the water, with toxicity apparent by a delay in moulting, even at <1.0 µg ASV labile copper·1−1. With all of these studies it must be realized that the addition of a complexing agent does not just reduce the level of toxic metal species but may change the speciation and hence the biological availability of a number of metals including those that are at low and limiting levels (e.g., Anderson & Morel, 1978). The biological availability of copper in sediments is controlled by the sediment particles themselves as well as by the organics in the interstitial water. This is through adsorption and complexation, either by the inorganic sedimentary particles or organic coatings on the particles (Lunz, 1972; Payne & Pickering, 1975; Pesch & Morgan, 1978; Sick, 1978). The biological effect and uptake of copper has also been found to be controlled by physical factors such as salinity (Jones, Royle & Murray, 1976; Stevens, 1977; Luoma, 1978b), temperature (e.g., Cairns, Buikema, Heath & Parker, 1978; Szeto & Nyberg, 1979), and pressure (Arcuri & Ehrlich, 1977). MacInnes & Calabrese (1979) examined the combined effects of salinity, temperature, and copper on embryos and early larvae of Crassostrea virginica and found that salinity had a greater effect on the embryos than temperature, at 0, 5, and 10 µg copper·1−1 but that temperature had a greater effect than salinity at 30, 60, and 90 µg copper·1−1. At 20 µg copper·1−1, temperature and salinity had equal effects suggesting a level of copper at which there is a changeover in the effect of the two variables. As the lowest levels of copper are the most likely to be found naturally, the interaction of copper and salinity are suggested as being the more important. Chemical factors, including metal interactions (e.g. Hutchinson, 1973; Braek, Jensen & Mohus, 1976; Yang & Ehrlich, 1976) have also been shown to be important, especially in estuarine conditions (e.g. Button & Dunker, 1971; Harding & Whitton, 1977). Melhuus, Seip & Myklestad (1978), however, point out that the interactions of metals depend on the nature of the actual chemical species. As the cupric ion has been demonstrated to be a biologically important species, (Swader & Chan, 1975; Sunda & Guillard, 1976; Sunda & Lewis, 1978), the importance of knowing the cupric ion concentration cannot be understated although accurate chemical measurement in the marine environment is difficult, if not impossible, at the present time.

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Biological factors also play a rôle in determining effects of copper. Larval stages of marine invertebrates generally have much lower tolerance to copper than do the adults (Wisely & Blick, 1967). Size has been shown to be important (Ayling, 1974; Boyden, 1977; Reeve et al., 1976, 1977a) although this may be a function of body volume and time to come into some sort of equilibrium with respect to metal levels. Seasonal variations in the level of organically bound copper and other metals (Morris, 1974) and even diel changes in “bound” metal (Johnson, 1978) have been linked to biological activity. This suggests a potential feedback system, changes in the biological availability being controlled by changes in the biology. Accurately predicting the biological effect of copper requires an understanding not only of metal speciation but also of the effect of the various metal species on the different developmental stages of the organism. With nutrients for example, it is possible to obtain some prediction of natural community structure based on nutrient physiology and competition (Kilham & Kilham, 1978). With speciation of copper in estuarine and marine environments it is possible, if one forgets about organic complexing agents, to obtain some estimates of ion-water and ion-ion interactions of the major constituents and how sea water affects the state of metal ions (e.g., Millero 1975, Westall, Zachary & Morel, 1976). These models can be used to determine the fate of copper added to the marine environment (e.g., Harrison, 1977c). When compared with toxicity data, they allow determination of metal species and toxic levels of copper (e.g., Howarth & Sprague, 1978). Mathematical models have also been developed to examine uptake of heavy metals in benthic algae (Seip, 1979) and other organisms. The relationship between the copper sensitivity of the calanoid copepod, Acartia tonsa, and the food ration has been described by a quadratic function which indicates that the log LC50 increases with increasing food ration to a point and then remains constant (Sosnowski, Germond & Gentile 1979). Both the concentration and species of metal in estuarine environments can be affected by the nature of the major industries in the region (e.g., Knauer, 1977; Mueller, 1977; Wong & Li, 1977; M.H.Wong, Chan & Choy, 1978). Metals tend to accumulate in the sediments, often with a change in the metal species due to chemical properties of the sediment (e.g., Whitfield & Lewis, 1976). In many estuaries dredging is commonplace for navigational purposes or landfill. This involves disturbing the sediments with their contained metals, often producing a change in the chemical properties of the metals and the potential of change in biological availability (Morton, 1977). One source of copper that has come into public focus in the past few years is copper tubing in large scale desalination plants and power plants, particularly nuclear power plants. Compton & Corcoran (1974) noted that single pass copper cooling systems add ≈ 1 µg copper·1−1 to coolant sea water. Chesher (1971) examined the biological impact of a large-scale desalination plant at Key West, Florida, and commented that ionic copper in the effluent was the most toxic feature. It was noted that the “start-up” after plant maintenance released large amounts of copper and thus caused more of a problem than under steady state conditions. Chesher also noted that extensive engineering changes were made to correct the corrosion in the pipes and to reduce copper discharge. Mandelli & McIlhenny (1971) examined the effects of copper in desalination brines and found that sea-water-brine dilutions containing 0·02 mg total copper·1−1 had an acute toxic effect on

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oyster larvae. This was in addition to the effect of changing temperature and salinity which, at 10% above the environmental level, was found to increase the level of the pathogenic fungus, Labyrinthomyxa marina, which is lethal to adult oysters. Mandelli (1975) found that of all the factors, the high copper content of the brines was the most critical. Romeril (1977) found that the brine effluent from a flash distillation desalination plant contained elevated levels of some heavy metals and increased metal concentrations were found in limpets and some species of Fucus near the outfall. Romeril & Davis (1976) examined trace metal levels in eels grown in power station cooling water, found that trace metals were not accumulated, and suggested that some mechanism prevents the accumulation of metals in eels. Hoss, Coston, Baptist & Engel (1974) attempted to determine the effects of temperature, copper, and chlorine on fish during simulated entrainment in power-plant condenser cooling systems. They used 1 mg copper·1−1 and found a synergistic effect of copper with temperature, that exposure of larval pinfish (Lagodon rhomboides) for 24 h affected their ability to withstand temperature shock. It was suggested from this work that “…protracted exposure to metal pollutants, such as copper, may reduce a fish’s ability to survive thermal shock”. Following the testing of a cooling system in the Diablo Canyon nuclear power plant in California, an estimated 1500 abalone (Haliotis rufescens and H.cracherodii) were killed in the discharge area (Martin, Stephenson & Martin, 1977). This was reported to have been from cooling system tests which were conducted after a period of closure during which non-circulating sea water was in contact with copper-nickel tubing in the condensing system. These authors noted that “it was apparent that large amounts of copper (Cu) dissociated from the tubing since the first pulse of discharge water contained 1,800 µg of Cu/liter. Concentrations rapidly decreased with flushing; however, even after 30 days, 20 µ Cu/liter were still found in the effluent waters versus 1 µg/liter entering the system.” As a result of the Diablo Canyon incident these authors examined copper toxicity in relation to abalone deaths observed in the power plant’s cooling waters. The work included examination of the effect of copper on both adults and larval stages although only seven adult specimens and the spawn from one male and female were used. Although the discussion states that 80 µg·1−1 can kill larvae and 50 µg·1−1 can kill adults, the authors did not relate this to concentrations within the bay where the abalone died nor did they consider the dilution of the effluent within the bay. The authors also found that copper accumulated in the gill tissues of both species at 56 µg copper·1−1 in the water and histopathological abnormalities in the gill tissues were found to occur at concentrations >32 µg·1−1. Later work (Yaffe, 1979) suggests that the most likely cause of death of abalone exposed to elevated copper concentrations is suffocation due to mucus smothering the gills. Harrison (1977b), in a quarterly progress report on chemical effluents in surface waters from nuclear power plants, found that copper at 1000 ng·ml−1 caused 95–100% inhibition of the spores of the alga Macrocystis and that the LC50 (48 h) of adult herring was 0.45 µg Cu·ml−1. Harrison (1977a,b) has also examined the effect of copper on the earlier life history stages of abalone. Her preliminary results indicate a decrease in growth with a copper concentration of ≈15 µg·1−1. One of the chemicals used in cleaning copper tubing in power plants and desalination

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plants is chlorine. (It is also used to reduce the coliform bacterial count in sewage.) Carpenter & Smith (1978) examined the effect of chlorine on the copper-complexing capacity of water in Biscayne Bay, Florida, and found that it has the potential for modifying organic-complexing agents resulting in a release of copper and an increase in copper toxicity. BIOASSAY ORGANISMS The sensitivity of certain organisms to heavy metals has led to their use as an assay for biologically available metal as well as a means of estimating biological effect. Chemical aspects of bioassay techniques for establishing water quality criteria are given in Lee (1973) while Phillips (1977) presents a review of the use of biological indicator organisms to monitor trace metal pollution in marine and estuarine environments. LaRoche (1972) discusses the value of short-term bioassays and Mancy & Allen (1977) have designed bioassays for determining the toxicity of heavy metals under controlled conditions. Bacteria have been used to measure the copper-chelation capacity of sea water (Gillespie & Vaccaro, 1978) as well as to measure cupric ion activity in sea water (Sunda, 1978). Sunda & Gillespie (1979) used 14C glucose incorporation by a marine bacterium as a response to cupric ion activity in sea water. They found that the relationship between inhibition of glucose incorporation and cupric ion activity fitted an equation derived from a molecular binding model:

where I is the rate of glucose incorporation in the presence of added copper, Imax is the rate in the absence of cupric ion inhibition, ACu is the cupric ion activity, and K* is a cellular inhibition site binding constant. Goulder, Blanchard, Sanderson & Wright (1979b) describe the use of the rate of glucose mineralization as a technique suitable for an assay of copper concentration while Bulich (1979) utilizes the light-producing response of luminescent bacteria for the rapid detection of toxic substances in water. Phytoplankton have been used as bioassay organisms (e.g., Erickson, Lackie & Maloney, 1970; Berland, Bonin, Maestrini & Pointier, 1973; Schmidt & Christensen, 1975; Davey, 1976b; Overnell, 1976; Hannan & Patouillet, 1977) as have benthic algae (Auerbach, Pruefer & Weiss, 1973; Haug, Sigurd & Sverre, 1974; Hostetter, 1976; Melhuus et al., 1978; Sivalingam, 1978). Eide, Jensen & Melsom (1979) used in situ cage cultures of phytoplankton to assess heavy metal pollution in two Norwegian fjords under conditions as close to normal as possible. They found that the growth of three species of phytoplankton (Skeletonema costatum, Phaeodactylum tricornutum, and Thalassiosira pseudonana) varied from species to species as well as from place to place, as did metal uptake of zinc, copper, lead, cadmium, and mercury. Barashkov & Kiristayeva (1977) used copper sulphate as an internal standard of toxicity with Chlorella in fresh water, and related the effect of other agents to the standard, expressed in units termed “toxes”, a tox being the ratio between the critical concentration of copper sulphate

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and the critical concentration of the substance being tested. The “critical concentration” is that which reduces the duplication time for Chlorella by over 5%. They also found that the critical concentration varied over the year, averaging 3 µg Cu2+·1−1. Christensen et al. (1979) used a freshwater green alga (Selenastrum capricornutum) and a euryhaline species (Chlorella stigmatophora) to examine the effects of manganese, copper, and lead in urban and agricultural run-off into a southern California marine basin. An algal bioassay test has been developed as an ecological test to define the effects of toxicants in the aquatic environment using Selenastrum capricornutum (Miller, Greene, Merwin & Shiroyama, 1978). Development of the test was with the understanding that a waste effluent may contain both organic and heavy metal components and that the effects of the heavy metals may be reduced by some organics. The results, in fresh water, suggest that growth inhibition is linear with an increase in zinc content of test waters but non-linear at higher levels of copper and cadmium. Russell (1979) used stenohaline isolates of Ectocarpus and other algae under low salinity conditions, to assay for copper tolerance of ship-fouling algae and Goodman, Newall & Russell (1976) used plasmolysis response in Ectocarpus spp. and rhizoid regeneration in Enteromorpha to estimate the effect of copper on ship-fouling algae. Hannan & Patouillet (1979) used an algal test to determine the toxicity of sediment elutriates in order to evaluate the effect of dredge spoils. From the assay results they suggest that toxicity is not directly related to total metal content of dredge spoils, and that analysis of total metal levels in dredge spoils is not applicable to the determination of the biological effect of the heavy metals in the spoils. Freshwater aquatic angiosperms have been examined as indicators of copper contamination (Ernst & Van der Werff, 1978). The results indicate specific responses by each species used, as a result of differences in rate of copper uptake and cellular localization of the metal. Not only are bioassay results subject to species specific responses but also to clonal effects whether in nature or in the laboratory (e.g., Fisher 1977). Genetic differences have been found in clonal cultures of two marine centric diatoms (Skeletonema costatum and Thalassiorsira pseudonana; Murphy 1978), differences which affect responses to environmental conditions (Murphy et al., 1978). In a study of the bioaccumulation of heavy metals by littoral and pelagic marine organisms Martin (1979) tested the concept that readily accessible intertidal marine invertebrates could provide useful indicators of heavy metal pollution. The tests indicated that serious problems existed in the application of the concept, problems similar to the species specific responses noted earlier for phytoplankton (Ernst & Werff, 1978). Stebbing & Pomroy (1978) used a hydroid, Hydra littoralis, to assess the effects of contaminants. Using the rate of a sexual reproduction they found that levels of 2–12 µg Cu2+·1−1 were inhibitory although, within the range of 0.5–2.5 µg Cu2+·1−1, the organism was able to counteract inhibitory effects. Stebbing (1979) advocates using bioassay techniques to evaluate the effects of individual pollutants in a multi-mixture medium, by chemically removing contaminants selectively from sea water. Evans (1977) utilized a coral community in a flow-through facility to examine the effects of copper while Karbe (1972) examined the effect of copper on a marine hydroid (Eirene viridula). In this latter study, threshold concentrations for acute effects were 0.03–0.06 mg·1−1, morphological changes and tissue re-organization were evident between 0.06 and 1 mg·1−1, and tissue

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disintegration occurred within a few hours at 3 mg copper·1−1. Guthrie, Davis, Cherry & Murray (1979) attempted to use in situ “microcosms” as an indication of the accumulation effect of heavy metals in run-off. The “microcosms” were clumps of oysters (Crassostrea virginica) which also contained barnacles (Balanus eburneus), blue crabs (Callinectes sapidus), clams (Rangia cuneata), and polychaetes (Nereis sp.). The idea is unique and attempts to provide an insight not only into accumulation of metals in a variety of organisms but also in organisms that interact. Copper levels in the environment and microcosms from two different places varied but, when averaged, suggested a species specificity in accumulation (water 1·31 mg·1−1, sediment 148.03 mg·kg−1 wet wt, barnacles 15.28 mg·kg−1, crabs 2.99 mg·kg−1, oysters 40.66 mg·kg−1, clams 22.75 mg·kg−1, and polychaetes 42.27 mg·kg−1 wet wt). Molluscs, particularly the mussel, Mytilus edulis, have also been used as indicator organisms (de Wolf, De Kock & Stam, 1972; Phillips, 1976a,b; Chaisemartin, 1977; Davenport & Manley, 1978; Harrison, 1978; Hummel & Speyer, 1978). M.edulis has become the major actor in the “Mussel Watch”, a programme designed to use selected organisms as an indication of pollution in the marine environment (Goldberg et al., 1978b; National Research Council, 1978). Harris, Fabris, Statham & Tawfik (1979) used M. edulis planulatus and other invertebrates as indicators of the effects of heavy metals in Western Port, Australia. Concentrations of cadmium, copper, manganese, and zinc, in M.edulis, were found to be a linear function of the length of the organism. Wood (1978) found that molluscs were useful for monitoring several metals “…because of their relatively long response time, which allows the tissues to integrate environmental concentrations”. Alexander et al. (1975) found that the digestive gland of the mussel M. californianus accumulated significantly higher levels of copper and other metals in areas where pollution input was significantly higher than natural input. A decline in the performance and physiological condition of M.edulis has also been found at increasing levels of pollutants, including copper, in Narragansett Bay, Rhode Island (Widdows, Phelps & Galloway, 1981). Few crustaceans have been used as bioassay organisms, possibly because of the difficulty of sorting out the effect of the cuticle in metal uptake (e.g., Subramanian, Yoshinari & D’Angeljan, 1974). The major exception is the brine shrimp, Artemia salina, which has been used for both fresh- and saltwater bioassay purposes because of the ease with which it can be obtained and grown. Alayse-Danet et al. (1979) for example, examined the effects of copper and zinc on amylase and trypsin activity. With trypsin, normal activity was disturbed within 72 h by a copper concentration of 2 mg·1−1 while amylase activity was disturbed within 24 h by the same concentration. Disturbances at the enzymatic level tended to occur more rapidly than decreases in growth rate and would thus allow a faster assay of conditions. Reeve et al. (1976, 1977a) and Reeve, Walter, Darcy & Ikeda (1977b) used natural copepod assemblages to examine sub-lethal responses to copper. Lewis & Ramnarine (1969) used the pre-feeding stages of Euchaeta japonica to examine the ability of natural systems to reduce the toxicity of copper while Sosnowski & Gentile (1978) examined the response of Acartia tonsa to cadmium, copper, and mercury through six generations. Both the reproducibility of the toxicological responses and the absence of demonstratable differences between field and laboratory populations suggest that A.tonsa is a species suitable for bioassay purposes.

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Sosnowski et al. (1979) did show, however, that the sensitivity of A.tonsa to copper was correlated with population density and food ration suggesting that differences in environmental conditions both within and between laboratory populations could produce a difference in bioassay results. Miller (1946) examined the effects of copper on the attachment and growth of the bryozoan Bugula neritina. His results showed that the critical copper concentrations that affect the various stages of the early life are: >0.3mg·1−1 kills larvae and inhibits the growth of adult forms, 0.2–0.3 mg·1−1 retards growth and prevents the formation of polyps, <0.2 mg·1−1 retards growth and polyp development; growth is inversely proportional to copper concentrations up to 0·3 mg.1−1. Timourian & Watchmaker (1977) developed an assay using the motility of sea urchin spermatozoa to study the effects of metal ions, including copper. Kobayashi, Nogami & Doi (1972) used sea urchin eggs to examine marine pollution in the Sea of Japan and found that the egg does not develop beyond the gastrula stage in water with high metal concentrations. Bougis, Corre & Etienne (1979) used the larvae of Paracentrotus lividus as a tool for assessment of water quality. Variability in the results, as well as survival under high levels of copper, however, reduces the usefulness of the organism for assay purposes. Fish have been used for routine assay of toxic effluents (e.g., Davis & Shand, 1978) with various routines being developed in an attempt to relate effluent toxicity to natural conditions (e.g., Stober, Dinnel, Wert & Nakatani, 1978). Morgan (1979) examined fish locomotor behaviour pattern as a monitoring tool for environmental bioassay purposes. Using opercular rhythms and activity responses, several agents were added to the flowthrough system. Opercular rhythm increased with decreasing copper as did activity with the two levels of copper used (0.1 and 1.0 mg·1−1). Schreck & Lorz (1978) developed a technique which allows the use of cortisol level in coho salmon to indicate stress with metals. In attempting to formulate any policy on copper and other metals in aquatic environments, it is important to recognize that one is dealing with organisms that respond differently from species to species as well as within each species, from one age group to the next. One must also be cognizant of the fact that physical and chemical factors influence the effect of the metal and that the interaction of these numerous factors produces an extremely complex situation. Winner, Scott Van Dyke, Caris & Farrel (1975) for example, point out that easily-derived diversity indices frequently give misleading evidence of stress, and that the nature of the stress and the nature of the biological community must be related. An additional problem in bioassay studies is the application of the laboratory results to field conditions. This stems from the problems associated with the requirements for controlled conditions in the laboratory and the use of levels of nutrients and metals not normally found in the field. It also unfortunately, comes from a lack of knowledge of the nature of factors controlling the biological availability of copper in natural environments. THE EFFECTS OF COPPER ON VARIOUS GROUPS OF ORGANISMS This subsection is intended to provide references dealing more with the organism than

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with physiological or ecological processes. Black, Hinton, Johnston & Sprague (1975) provide an annotated list of copper concentrations found harmful to aquatic organisms. This is a review of selected literature with the original intent to provide. “…succinct overall conclusions on the thresholds for effects of copper”. Gross et al. (1971) surveyed the marine waste deposits in the New York metropolitan area and provide copper levels and species composition in polluted regions. Pesando & Aubert (1975) examined the effect of pollutants on a variety of organisms, finding that the level of copper used (0.62 mg·1−1) did not have any measurable effect. Beers, Stewart & Hoskins (1977), using the CEPEX bags, examined the dynamics of micro-zooplankton populations treated with copper. Bacteria Patrick & Loutit (1976) examined the passage of metals in effluents, through bacteria to higher organisms, and suggested that small amounts of metal are concentrated by the bacteria and thus introduced into the food chain. Patrick & Loutit (1977) also found that bacterial epiphytes on the surface of a freshwater plant (Alisma plantago-aquatica) collected from a polluted river, contributed to the total metal concentration of the plant. The removal of the epiphytes resulted in a 30–35% reduction in the concentration of copper in the plant as well as major losses in other metals. They found that the concentration of the metal in the epiphytes remained about the same in unpolluted water but that the number of epiphytes was higher in the polluted water. They inferred that a variation in the number of epiphytes gave the increase in metal concentration seen in the plant in polluted waters. G.E. Jones et al. (1976) found that Arthrobacter marinus and Pseudomonas cuprodurans assimilated trace metals against a gradient in the order: . They also found that “as the chelating capacity of the nutrients in the medium was almost exceeded, the Zn assimilation was accelerated”. Albright, Wentworth & Wilson (1972) found the order of sensitivity of a culture of mixed bacteria to sublethal concentrations of metal to be others. Jones (1967) showed that Escherichia coli grew poorly in sea water but that there was a marked improvement in growth after autoclaving the sea water or adding a chelating agent. He suggested that it was the combined metals present in sea water which were toxic as the levels of copper alone were not sufficient to produce the observed toxic effect. The toxic effect of metal ions in sea water was stated as being the principal factor in the death of freshwater bacteria in sea water (Jones & Cobet 1975). Den Dorren De Jong (1971) showed that 10−4M CuSO4·5H2O inhibited the growth of Aztobacter, Hata (1960) found that the growth and activity of marine sulphate-reducing bacteria was inhibited by the same concentration of copper sulphate, and Singleton & Guthrie (1977) found that both copper and mercury additions caused a reduction in species diversity of bacteria. Yang & Ehrlich (1976) found that the effects of four metals (Mn, Ni, Cu, Co) on deep-sea bacteria could be controlled by the nature of the medium used. Vaccaro, Azam & Hodson (1976) noted that bacterial heterotrophs exposed to 10 and 50 µg Cu2+·1−1 in CEPEX enclosures initially increased in activity, presumably due to organic carbon from copper-sensitive plankton in the enclosure. With time, the

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surviving bacteria developed an increased copper tolerance. Mitchell (1974) showed that sublethal levels of copper and other pollutants upset the microbiological balance in reef corals, that the organic material released by the corals caused an increase in the bacterial population that helped to cause the death of the corals. Gillespie & Vaccaro (1978) isolated a gram-negative, motile, rod-shaped organism that they used as a bacterial bioassay for measuring the copper-chelation capacity of sea water. In two rather similar papers Goulder et al. (1979b) and Goulder, Blanchard, Metcalf & Wright (1979a) demonstrated that bacteria can be used in the recognition of pollution stress from metal refinery effluent in estuaries. Waksman, Johnston & Carey (1943) found that bacteria had an affinity for copper in the sea-water surface film while Kurata & Yoshida (1978) examined the heavy metal tolerance of 219 bacterial strains from sea water and sediments and found that those from sediments tended to tolerate higher levels of copper than those from the water. Austin, Allen, Mills & Colwell (1977) examined the taxonomic relationships between heavy metal tolerant estuarine bacteria. Sunda & Gillespie (1979) used a bacterium to estimate cupric ion activity in sea water while Zevenhuizen et al. (1979) found that the copper-sensitive bacterium Klebsiella aerogenes was inhibited in growth and survival by copper ion concentrations of 10−8– 10−6M. Fungi Avakyan & Rabotnova (1966) determined the level of copper toxic to the non-marine Torula (=Candida) utilis to be >30mg·1−1 at pH 5 and >20mg·1−1 at pH 7. The yeast was shown to be unable to develop a tolerance to higher levels of copper, even after repeated subculture in a glycerin medium. A number of organic acids were tested to determine their ability to reduce the toxicity of copper to the yeast and one of the authors (Avakyan, 1971) found that toxic effects were reduced when a copper atom entered into a coordinately saturated complex, while co-ordinately unsaturated complexes retained some of their toxic properties. Zhirova, Ivanova & Grachev (1977), working with the nonmarine yeast Saccharomyces carlsbergensis, found that copper had no appreciable effect on the formation of acids and ethanol metabolites. Button & Dunker (1971) found that the inhibitory effects of copper toward marine yeasts were prevented by phosphate. Copper was found to be inhibitory in phosphate-limited systems only under conditions of manganese deficiency suggesting a rather complex link between metals and nutrients. Keyhani (1973) found that when subjected to higher concentrations of copper, yeast mitochondria become enlarged and the cisternae were less well developed. Schneider (1972) found 10−2M copper to be acutely toxic to a marine fungus but that the toxicity was controlled by salinity. Marine Plants Copper has been shown to be effective in reducing the production of algal blooms (Muehlberger, 1969; Sawyer, 1971; Dashora & Gupta, 1978) although blooms may also be copper deficient (Telitchenko et al., 1970). Copper has been found to inhibit the growth of several dinoflagellates (Saifullah, 1978) and Anderson & Morel (1978, 1979)

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imply that organic chelation of copper may be necessary before the dinoflagellate Gonyaulax tamarensis can successfully compete with other phytoplankton in coastal waters. They suggest (1978) that natural levels of the cupric ion may be toxic, something that Barber & Huntsman (1978) also suggest in an examination of phytoplankton growth in deep ocean sea water. Copper complexation has also been suggested as important to diatoms and organic agents produced by certain phytoplankton may play an important rôle in species succession (Lewis 1977; Gnassia-Barelli, Romeo, Laumond & Pesando, 1978). The nature of copper complexes has been shown to affect toxicity of copper to algae although interspecific differences occur in the response to the complexes (Khobot’yev, Kapov, Rukhadze & Turunina, 1975) as well as to copper itself (Mandelli, 1969). Bentley-Mowat & Reid (1977), for example, found differences in the response of several phytoplankton species to copper in a growth medium (Droop Medium S88) as did Kling (1974) with the freshwater alga Gracilaria verrucosa in another growth medium (modified Fries ASP6F medium). Rosko & Rachlin (1975) found that in both chelating and non-chelating media the order of toxicity for several metals to the diatom Nitzschia closterium was . Copper has also been found to affect the dominance and diversity of algae (Thomas & Seibert 1977). Knauer & Martin (1973) found that phytoplankton uptake of copper appeared to have little effect on the concentration of dissolved copper in Monterey Bay, California. Jennett, Hassett & Smith (1977) tested several freshwater green and blue-breen algae to determine if metals could be concentrated from the natural environment. They comment that “…adsorption of metals almost certainly depends on the production of extracellular products by algae such as heteropolysaccharides, glycolic acid, and glycoproteins”. The possibility that many of these agents are released into the water (e.g., Gnassia-Barelli et al., 1978) suggests that the ability to concentrate metals from the natural environment is dependent upon the needs of the cell and the ability to take up organometallic compounds. Spencer & Brewer (1969) found a relatively low incorporation of copper, zinc, and nickel into phytoplankton in the euphotic zone although Weigel (1977) found a significant correlation (1% level) between particulate copper and chlorophyll in the photic zone of the Baltic Sea but not in deep water. Danielsson (1980) examined the relationship between copper and silica in diatom frustules and found a total Cu/Si molar ratio of 2.4×10−5 in Indian Ocean waters although they point out the possibility that the values may have been biased by contamination during the sampling programme. Young & Lisk (1972) found that a combination of copper and silver was more effective in controlling freshwater algae than copper alone. They point out, however, that the combined cost and environmental effect limits the usefulness of the silver. Harrison (1977c) found that the spores of Macrocystis could be inhibited by copper (1000 ng·1−1). Resistance to copper is not uncommon in macroalgae (e.g., Russell & Morris 1970). In both phytoplankton and macroalgae, the biological effect of the metal is controlled not only by the state of the organism but also by the state of the metal (e.g., Pakalne, Nollendorf & Upitis 1970; Pieterse, Bhalla & Sabharwal, 1970; Steeman Nielsen & Wium-Anderson, 1970; Upitis, Pakalne & Nollendorf, 1971; Davey, Morgan & Erickson, 1973). The work of Murphy (1978), Murphy et al. (1978), and Murphy & Belastock (1980), on factors producing variability within phytoplankton organisms, suggests that

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genetic change may occur in phytoplankton cultures with time and that this potential for change, in the natural environment, may produce physiologically different populations of a species. The result is an intraspecific difference in response to metal. Variation in metal content of macroalgae has been found to be a function of the proximity of the organism to the source of the metal (Foster, 1976). Seeliger & Edwards (1977) examined the concentration factors of copper and lead in sea water and benthic algae in the New York metropolitan area, finding a concentration factor of 18.4×104 for copper in some benthic algae and a strong correlation of copper content to copper in the water. Seeliger & Edwards (1979) also examined the fate of biologically accumulated copper in growing and decomposing thalli of two benthic red marine algae (Ceramium pedicellatum and Neogardhiella baileyi). Concentration factors for copper in the thalli were 4540 to 6864 with plants grown in organic-free sea water with added copper. Subsequent measurement indicated that up to 22% of the copper was bound to dissolved organics released by the thalli while only 3–10% was associated with dissolved inorganics. Copper in decomposed thalli was present in both dissolved and detrital fractions. Rodgers, Cherry & Guthrie (1978) examined cycling of elements, including copper, in the freshwater duckweed (Lemna perpusilla) and, like Seeliger & Edwards (1979), suggest that the bioaccumulation of metals by plants forms a mechanism for introducing metals into food chains. North (1964) found that at high levels of copper the metal was first taken up by the blades of the kelp Macrocystis pyrifera, that the blades exuded a yellow substance, and that the plant had a definite “toxicity syndrome”. North (1964) and Clendenning & North (1960) both found that copper levels of 0.1 mg·1−1 produced a noticeable reduction in photosynthesis in M.pyrifera. Wolfe, Thayer & Adams (1976) found that the manganese, iron, and copper incorporated into the annual production of the eelgrass Zostera marina “…constituted significant fractions of the amounts of these metals contained in one year’s accumulation of sediments”. Windom (1975) found that the average annual uptake of copper by Spartina alterniflora leaves and stalks in a salt marsh estuary was 103 kg or about 3% of the total river input into the estuary. Myklestad, Eide & Melsom (1978) found that transplanting Ascophyllum nodosum from an area of high concentration of metal to an area of low concentration caused new plant material to be lower in metal concentration than the older material. Although copper was not considered because it was at elevated levels in both areas, results with the other metals (Zn, Pb, Cd, Hg) suggested that once the metal was incorporated into the tissue, it tended to be retained. Protozoa Hanna & Lilly (1974) found that the marine ciliate Uronema marinum exhibited no requirement for Fe, Zn, Mn, or Cu in a chemically defined medium although high levels of the metals were found to be toxic. Ruthven & Cairns (1973) examined the response of freshwater protozoan communities to metals. They showed that while a particular species may be tolerant to relatively high levels of copper they are not necessarily tolerant to other toxicants. Cairns & Dickson (1970) showed that, like most other groups of organisms, high concentrations of copper will lead to survival only of the most tolerant species.

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Invertebrates Wisely & Blick (1967) determined the effect of copper on the larval stages of a number of invertebrates and noted that levels of copper toxic to the larval stages are generally less than those levels found toxic for the adult organisms. Mitchell (1974) and Chet & Mitchell (1975) examined the stress-induced attack of corals by marine bacteria using petroleum hydrocarbons and heavy metals (including copper) as stressing agents. They found that the pollutants stimulated excessive extracellular production of a polymer which attracted motile, gram-negative pseudomonads that predated on the tissue of the corals. Emerson (in Soule & Oguri, 1976) examined heavy metal uptake of two polychaetous annelids from resuspended sediments to determine the potential biological effect of heavy metals in dredged sediments. The results showed no increase in heavy metal levels in the tissues which Emerson suggested was due to the “scavenging” ability of suspended sediments. Emerson did not consider, however, the possible complexing ability of sediment elutriate (e.g., Lewis, Whitfield & Ramnarine, 1973). Reish & Carr (1978) examined the effects of heavy metals on the survival, reproduction, development, and life cycles of two species of polychaetous annelids in southern California. They found a significant reduction in the production of offspring at copper (and other metal) levels less than the 96-h LC50 level. The 96-h LC50 level of copper for the two species was within the range of concentration found in a sewage effluent discharge, suggesting that some effect from copper could be occurring in the area of the discharge. Bryan (1976) examined heavy metal tolerance of Nereis diversicolor as well as other marine organisms. In Nereis Bryan found a close relationship between the rates of absorption of metals and their acute toxicity, metals such as copper and silver being much more toxic than zinc. McLusky & Phillips (1975) suggested that, in Phyllodoce maculata, the rate of uptake of copper may be the lethal factor rather than the amount of copper accumulated. MacInnes & Calabrese (1979) examined the combined effects of temperature, salinity, and copper on the embryos and early larvae of Crassostrea virginica and noted a change in the effect of temperature and salinity with increasing levels of copper. Sankaranarayanan, Purushan & Rao (1978) noted seasonal changes in the concentration of copper, zinc, and iron in C.madrasensis from the Cochin backwaters in India which they related to monsoonal changes in freshwater run-off from areas receiving industrial and domestic wastes. Mandelli (1975) noted that the copper content of brines from desalination plants was sufficient to affect adversely juvenile and adult C.virginica. The effects of copper on clams have been examined by a number of individuals and groups (e.g. Orton, 1923). Eisler (1977b) found that a mixture of metal salts (including copper) which approximated the highest measured levels in the sediments of Narragansett Bay, Rhode Island killed Mya arenaria. McGreer (1979a,b) found no evidence that copper in sediments around a sewage outfall was toxic to the adult of Macoma balthica but suggested that it may be a controlling factor at the time of larval settlement. Marks (1938) showed that 0.1–0.2 mg copper·1−1 was toxic to several species of molluscs. Kumaraguru & Ramamoorthi (1978) gave 96-h LC50 values of copper for three estuarine bivalves, with values of 60 µg Cu·1−1 for Anadara granosa, 72µg Cu·1−1 for Meretrix

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casta, and 88 µg Cu·1−1 for Crassostrea madrasensis. Olson & Harrel (1973) found that the toxicity of copper to an estuarine clam increased with a decrease in salinity. Floch, Deschiens & LeCorroller (1964) examined the molluscicidal effects of a variety of compounds in fresh water to determine the most appropriate compound to eliminate bilharzia vectors. Lakshmanan & Krishnan Nambisan (1977) found that Villorita cyprinoides var. cochinensis accumulated copper from the surrounding medium, that concentrations of 0.5 mg·1−1 were sublethal, and that the effect of a chelating agent, EDTA, was to reduce toxicity. Geldiay & Uysal (1976) noted highest copper levels in the mantle and siphon (10.4µg·g−1 dry weight) and lowest in the muscle tissue (1.7 µg·g−1) of Tapes decussatus in the Izmir Bay area of Turkey. Larsen (1979) found that copper varied significantly with the age of Mercenaria mercenaria in the lower Chesapeake Bay region. One of the considerations that must be made when examining any data concerning toxic concentrations is the state of the metal and the effect of organics and particulates on that state. Floch et al. (1964), Paulin, Camey & Pereira (1963), and Frick, Ritchie, Fox & Jimenez (1964) found Cu2O to be highly toxic while Floch et al. (1964) found less toxicity with CuCl2 and elemental copper. Huguenin (1977) found that the oyster drill, Urosalpinx cinerea, would not cross a strip of metallic copper and suggested that it would be a useful technique to reduce predation on oysters. Spronk, Brinkman, Van Hoek & Knook (1971) suggested that copper inhibits Na+– K+–ATPase in the pond snail Lymnaea stagnalis which results in a net increase in the water content of the kidney and less effective sodium exchange. In a study of the distribution of major and minor elements in marine animals, Segar, Collins & Riley (1971) found a higher concentration of copper in the soft parts than the shells of 11 species of molluscs from the Irish Sea as well as one freshwater species. Bryan & Uysal (1978) found that only 30–40% of the total silver and copper in the burrowing bivalve Scrobicularia plana was in the digestive gland and from this implied that uptake from solution could be as important or more important than uptake from the sediments that are ingested. Martin & Sparks (1971) found, however, that the first tissue to show an increase in copper in Corbicula fluminae was the digestive diverticulum, then the gills, followed by the mantle epithelium. Sheppard (1977) found changes in several major cations in animal tissue along some pollution gradients which are suggested to be associated with the physiological stress of the organisms in a polluted environment. Scott & Major (1972), working with Mytilus edulis, found that copper caused respiratory and cardiovascular depression. Brown (1972) obtained similar results and also showed that sublethal levels of copper inhibited ciliary activity. Martin et al. (1975) found that copper caused a decrease in the number of byssal threads in M.edulis and Manley & Davenport (1979) noted shell closure of M.edulis to be faster than that of Modiolus modiolus and M.demissus after addition of various copper concentrations. The authors also suggest that closure may be initiated by the organism’s response to a specific form of the metal. Davenport (1977) found continuous application of copper damaging to Mytilus edulis and related it to the position of the organism in the tidal zone and the activity of the organism as affected by salinity and tide. Hueck (1975) found that mussels accumulated copper although they were able to eliminate some of the metal. Harris et al. (1979) noted that the concentration of several metals including copper, in M.edulis

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planulatus, varied with the length of the organism while Fowler & Oregioni (1976) found metal levels in mussels to be highest in samples from ports and areas of river discharge in the Mediterranean. Collinson & Rees (1979) found high mortality of mussels during dredging operations in the Gulf of La Spezia in Italy and attributed this partly to copper. Alexander & Young (1976) found M.californianus a useful indicator of anthropogenic inputs of several trace metals off southern California and noted highest concentrations in the digestive gland. They pointed out that copper, chromium, and silver in the mussel are closely related to “urban point sources”. Majori & Petronio (1973) found accumulation of copper in M. galloprovincialis and provided calculated concentration factors for a number of metals. The effects of copper on crustaceans have been studied both from the standpoint of the copper-containing blood pigment haemocyanin as well as from the deleterious effects. Dow et al. (1975) comment that “of those metals which have been evaluated copper causes the highest rate of mortality among lobsters. Although natural sea water contains this metal in measurable amounts, any appreciable increase in the copper content of water in which lobsters are held usually causes mortality of the animals.” Neff & Anderson (1977) examined the effects of copper on moulting and growth of juvenile blue crabs Callinectes similis, using 500 µg copper as CuSO4·1−1. They found that this level was acutely toxic to the megalops and juvenile stages, recognized that the chemical form of copper was important in examining toxicity, and found that the majority of deaths were during or immediately after a moult. Gibson, Thatcher & Apts (1975) examined the effects of temperature, chlorine, and copper on Pandalus danae and found that 0.041 mg Cu·1−1 effectively retarded growth of 1–2-gm individuals at 16°C over the period of one month. Holm-Jensen (1948) examined the effect of heavy metals on osmotic regulation in Daphnia magna and found that copper toxicity was decreased by increasing salinity. Copper uptake was also found to be a function of copper concentration and salinity. In fresh water, Biesinger & Christensen (1972) found that different levels of copper caused reproductive impairment, decrease in growth, protein content, and glutamic oxaloacetate transaminase activity in Daphnia. These authors also examined the toxicity of various forms of copper to D.magna, finding that CuCl2 was less toxic than CuSO2 and that nitrilotriacetic acid (NTA) complexes of both copper and zinc were relatively non-toxic. (See Sanchez & Lee, 1973, for a discussion of some of the characteristics of NTA.) Corner & Sparrow (1956) found that copper caused a depression in the respiratory rate of Artemia salina but had no effect on mortality at the concentrations that they used. They also found that A.salina could be sensitized to mercury salts but not to copper salts by a sublethal dose of copper. Brown & Ahsanullah (1971) examined the effect of sublethal concentrations of heavy metals on A.salina and the polychaete worm Ophryotrocha labronica and found a difference in susceptibility of the two species as well as noting that the larvae were more susceptible than the adults. At 0.1 mg·1−1 the growth rate of Ophryotrocha was decreased significantly while at 1.0 mg·1−1 the growth of Artemia was not changed. In sea-water-acclimated Artemia, Saliba & Krzyz (1976) found that copper toxicity depended on the particular form of the metal as did growth inhibition and acquisition of copper tolerance. Connor (1972) examined the sensitivity of several organisms to copper and the results indicated that larval stages exhibited greater

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sensitivity than the adults. Hubschman (1966) found that, at molar levels of copper, the toxic action in Orconectes rusticus was the result of protein coagulation. Millimolar levels of copper inhibited respiratory enzymes and succinate metabolism and micromolar levels had a chronic effect on cell maintenance. Chaisemartin (1973) found that, of the members of the family Astacidae that were examined, members of the genus Orconectes tended to be the most tolerant. Chaisemartin also found considerable interspecific differences in response to copper and noted that the highest mortality rates occurred during ecdysis and during the first year of life. Barnes & Stanbury (1948) examined the effects of copper and mercury salts on Nitocra spinipes. In the range 0.026–26.0 mg Cu·1−1, a tenfold increase in copper lead only to an additional 11–21% of the animals being killed. It thus appeared that either the organism was very tolerant or that the effect of copper was modified by some environmental condition. Barnes & Stanbury also showed that mixtures of copper and mercury salts were much more toxic than either of the metals separately. They suggested that the action of copper on the organism was slow and that the addition of mercury impaired respiratory and excretory function to a point where the effectiveness of copper as a toxicant was increased. D’Agostino & Finney (1974) found that copper and cadmium inhibited the growth and development of the F-2 generation of Tigriopus japonicus, and that the metals together were much more toxic than separately. McConaugha (in Soule & Oguri 1976) found a reduction in the survival of Acartia tonsa, and Tisbe sp. in 0·45 µmfiltered water in which bottom sediments had been suspended. The suggestion was that the reduction was due to heavy metals in the sediments. The crab Pachygrapsus crassipes, exposed to the filtrate for seven days, however, exhibited no change in metal levels in gill tissue. Polikarpov, Oregioni, Parchevskaya & Benayoun (1979) measured the body burden of chromium, copper, cadmium, and lead in Anomalocera patersoni collected in the Mediterranean Sea. The copper content did not differ significantly between males and females and tended to follow a logarithmic normal distribution. MoraitouApostolopoulou & Verripoulos (1979) examined the effects of sub-lethal concentrations of copper on Acartia clausi and found some evidence that individuals from regions of higher pollution were better able to tolerate high levels of copper than those from regions with little pollution. This suggests that acclimation could occur, something that Murphy & Belastock (1980) found with Thalassiosira pseudonana. The results of MoraitouApostolopoulou & Verriopoulos (1979) should be examined in the light of the findings of Sosnowski & Gentile (1978), that the toxicological responses of six generations of Acartia tonsa to cadmium, copper, and mercury are similar. This suggests that if physiological populations of A.clausi do occur they have taken a number of generations to develop. Pyefinch & Mott (1948) showed that larval stages of several barnacles had different sensitivities to copper, the nauplius stages being much more sensitive than the cyprids. These authors also demonstrated an interspecific difference in response and an effect of salinity on the toxicity of copper. Bernard & Lane (1961) found that metal accumulation did not increase with increasing copper in the cyprids before settlement. They also suggested that copper was important in the metabolic change which occurs at metamorphosis and that, in the adult barnacle, the toxic action of copper was attributable

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to its interference with normal respiratory processes. Boulton, Huggins & Munday (1971) examined the toxic effect of a variety of organometallic compounds on Elminius modestus, with the intent of improving anti-fouling paints. They found that copper citrate disturbs pyruvate metabolism and interferes with the flow of carbon into the citric acid cycle and suggested that the most likely site of inhibition was pyruvate dehydrogenase. Hunter (1949) found that a copper concentration of ≈50 mg·1−1 was toxic to the amphipod Marinogammarus marinus, especially under conditions of low salinity and low oxygen concentration. Copper also had a marked effect on animals which were under stress from mercury. Based partly on these findings Hunter suggested that the action of mercury was direct while that of copper was indirect. The effect of copper on macro-zooplankton populations was examined in the CEPEX enclosures by Gibson & Grice (1977). They found a severe reduction in zooplankton abundance in both the copper enriched and control containers and suggested that the decline was due to grazing which made it difficult to assess the effect of copper. At a nominal concentration of 50 µg Cu·1−1 the abundance of two common calanoid copepods (Pseudocalanus sp. and Acartia longiremis) was reduced to 50% of the original levels 3– 3.5 times more rapidly than in the controls. The results with lower concentrations (5–10 µg Cu·1−1) were more variable and less significant. Fowler (1977) measured trace elements in particulate materials derived from zooplankton. Faecal pellets contained the highest levels with lesser amounts in moults and eggs. The concentration of copper in euphausiid moults was not significantly higher than in whole individuals. Fowler calculated the trace element flux rate contribution by euphausiids (µg·kg euphausiid−1·day−1) for faecal pellets (Cu=8600) and moults (Cu=320) and pointed out the importance of this in vertical movement of particulate copper in the oceans. Lillie (1921) examined the effects of copper on the fertilization reaction of Arbacia. Inhibition of the fertilization reaction occurred at copper levels of 0.4 mg·1−1. Copper ions affected the activation of the process but not the reaction and the inhibitory effect was reversible. It was suggested that the inhibitory effect of copper was the result of its combination with the fertilizing of the egg. Lillie (1921) also showed that there is a distinct difference between the action of copper and mercury. Allee, Finkel & Garner (1941) found that the rate of cleavage in Arbacia was proportional to the concentration of copper present. Finkel, Allee & Garner (1942) found similar results and further suggested that copper acted directly in stimulating the rate of cleavage at concentrations of 10−13– 10−7M. Bougis (1965) and Bougis & Corre (1974) examined the effect of copper on the pluteus of Paracentrotus lividus. Grave (1941) found an acceleration of development in ascidian larvae which was suggested to be the result of the action of copper on an inhibiting enzyme. Glaser & Anslow (1949), on the other hand, attributed the acceleration due to copper to its ability to activate certain enzymes. Vertebrates Metal levels in marine fishes are of interest because of the importance of the fishes as a source of food. Majori, Nedoclan, Modonutti & Daris (1978) found geographic differences in the concentration of copper in commercially important marine fishes in the

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northern Adriatic Sea. MacKay, Kazaco, Williams & Leedow (1975) reported relatively low concentrations of cadmium, copper, lead, and arsenic in muscle tissue of black marlin taken off northeastern Australia with an average copper level of 0.4 mg·kg−1 wet weight in the muscle tissue and 4.6 mg·kg−1 wet weight in liver. McDermott, Alexander, Young & Mearns (1976) examined the concentration of a number of trace metals in Dover sole (Microstomus pacificus) taken near waste-water discharges in California and found a 20-fold copper enrichment of the sediments in the outfall areas but no copper enrichment of flesh, gonads or liver. This trend was also found in “finfish” from the New York Bight and Long Island Sound by Greig & Wenzloff (1977). Levels of individual metals also varied little between species and, within a species, between specimens from the heavily polluted New York area and the less polluted open North Atlantic. These studies suggest a capability in fish of eliminating excess metals, possibly through a metallothionein-type system (e.g., Olafson & Thompson 1974). A study by Coombs (1976) showed that exposure of plaice to sublethal concentrations of cadmium produced significant changes in the zinc and copper distributions in the tissues. This could very easily be due to the stability constants exhibited by a metallothionein-like agent for the various metals in combination with the level of the metal and the ability of the organism to produce the complexing agent. Rice & Harrison (1978b) used a flow-through bioassay system to examine the copper sensitivity of Clupea harengus, during its early life. Significant embryo mortalities occurred at 35 µg Cu·1−1 although the time of the exposure relative to the age of the embryo produced differences in the mortality. The effects of sublethal concentrations of pollutants (copper, DDT, parathion) on fish behaviour formed the basis for a study by Kleerekoper, Matis & Rand (1975). Zahuranec (1978) reviewed the shark repellant screening tests, of various chemical elements, that have been done for the U.S. Office of Naval Research. A mixture of 20% copper acetate and 80% nigrosine dye proved effective, sometimes nearly 100% effective, in keeping sharks from feeding, even when they were already actively feeding on trash fish shovelled off the deck of a shrimp boat. Some of the effects of copper on freshwater fishes are of importance in considering the overall effect of the metal on marine fishes. Jones (1935, 1937, 1940, 1941, 1942) examined the effects of a variety of metals on several species of freshwater fishes and found that the toxic action of the metallic cation was only evident in hypertonic solutions. The lethal limit for copper to Gastrosteus aculeatus was ≈60 µg·1−1 (Jones, 1938) although the toxic effects of certain metals could be reduced by calcium salts (i.e., increasing the hardness of the water). Kleerkoper, Westlake, Matis & Ginsler (1972) showed that goldfish orientation was affected by a concentration gradient of 11–17 µg Cu·1−1 and Sprague (1964a) found that young Atlantic salmon will avoid copper and copper-zinc solutions. The threshold levels for avoidance were 2·3 mg copper·1−1 and, in the presence of 6.1 mg zinc·1−1, 0.42 mg copper·1−1. These were, respectively, 10 and 7% of the incipient lethal levels. There was also no evidence of the fish becoming acclimatized or sensitized by previous treatments. (The greater toxicity of the combined copper-zinc solutions than either of the metals alone is similar to the results found by Hunter, 1949, for crustaceans.) Sprague, Elson & Saunders (1965) showed that when copper and zinc were present in the Miramichi River, downstream moving salmon avoided the regions where the levels of the metals were

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high. This suggests that one of the important toxic actions may be indirect, through the creation of an avoidance reaction (Saunders & Sprague, 1967). The addition of nitrilotriacetic acid to water with high levels of copper and zinc was found to slow the avoidance reaction of fish (Anonymous, 1968) suggesting that natural or synthetic complexing agents could change the reaction of the fish. Koeller & Parsons (1977) found no direct effect of copper (2.5 µg·1−1) on the growth or survival of young salmonids in the CEPEX enclosures and suggested that the effect of the copper would most likely be through the effect on the diversity of the food available. Sprague (1964b) determined the lethal levels of copper to adult Atlantic salmon to be 48 µg·1−1. Sreenivasan & Sounder Raj (1963) found that increasing the hardness of the medium in which the organism was held caused a reduction in the toxicity of zinc. (This result is not unexpected because of the chemical shift in the inorganic metal species and the reduction in the concentration of the ionic form.) Sprague & Ramsay (1965) reported lethal levels of copper for juvenile Atlantic salmon to be 32 µg·1−1 in water with a hardness of 12 mg CaCO3·1−1. The difference between the results in Sprague (1964b, above) and Sprague & Ramsey (1965) is probably due to the difference in hardness of the waters. Fitzgerald & Faust (1963) showed that the toxicity of copper to minnows was dependent on pH and could be reduced by the chelation of copper with citric acid. Samylin (1966, 1967) showed that copper naphthenate was lethal to Atlantic salmon at concentrations ≥3000 mg·1−1. Tabata (1969a,b,c), Nishikawa & Tabata (1969), and Tabata & Nishikawa (1969) provide further documentation on the effects of copper and other metals to fish. Tabata (1969a) showed that heavy metal toxicity could be reduced by increasing the water hardness and also that metallic cations had an antagonistic effect on the toxic activity of heavy metals. Nishikawa & Tabata (1969) showed that the metal, when in complexed form, is less toxic than the ionic form. Tabata (1969b) suggested that the quality of environmental water can be predicted from its heavy metal content and its hardness but only after determining the relationship between the metal and the organism and assuming that the relationship remains constant. Tabata & Nishikawa (1969) found that the addition of EDTA caused a reduction in the toxic effects of high metal concentrations in industrial wastes. Wilson (1972) examined the possibility of producing a simple model for predicting the copper toxicity of waters receiving an input of copper. He concluded that variations in the hardness of the water may be sufficiently great during certain periods of the year to make long term estimations of toxicity impractical. The effects of other metals in conjunction with copper (e.g., Sprague, 1964b; Sprague & Ramsay, 1965) would also make predictability difficult, especially when the effects of metal pairs were not known. Even if these problems were solved one would still have to consider the wide variety of organic compounds that have been shown to complex dissolved copper (Nishikawa & Tabata, 1969; Lewis et al., 1972, 1973). Utiger (1968) showed that copper inhibited the development of secondary sexual characteristics of minnows. This was attributed to the effect of copper on the enzymes responsible for the hormone producing the characteristics. Jackim, Hamlin & Sonis (1970) showed that five liver enzymes of the killifish Fundulus heteroclitus have markedly different activities in fish which have survived exposure to the median lethal dose of copper for 96 h when compared with

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unexposed fish. It was suggested that a biochemical autopsy could be a useful tool in determining the sublethal effects of metal cations. Ozaki, Uematsu & Tanaka (1970) found that sub-acute levels of copper compounds acted on the mucus epithelium of carp. Hazel & Meith (1970) showed, with a flowing water bioassay, that the eggs of the king salmon (Oncorhynchus tshawytscha) are more resistant to the toxic effects of copper than were the fry. Copper concentrations of 0.08 mg·1−1 did not affect the hatching success of the eggs although concentrations of 0.04 mg·1−1 were acutely toxic to fry and concentrations of 0.02 mg·1−1 caused increased mortality and inhibited growth. Blaxter (1977) noted the effect of copper on the eggs and larvae of Pleuronectes platessa and Clupea harengus and Voronina & Gorkin (1978) examined the impact of copper on the early period of development in Tilapia in the marine environment. McKim, Christensen & Hunt (1970) showed that short term exposure to various levels of copper caused an increase in the red blood cell count, haematocrit, and haemoglobin, with decrease in plasma chloride and osmolality in the trout Salvelinus fontenalis. After long term exposure there was only a decrease in plasma oxaloacetate transaminase at the highest levels used. This suggested that the changes noted for the short term study were only transient features of the initial responses. Gardner & LaRoche (1973) found cellular changes attributable to copper in chemoreceptors of the lateral line canals and head of two estuarine teleost fishes (Fundulus heteroclitus and Menidia menidia). They also noted that the fry of Fundulus were more sensitive to copper than the adult or the zygote and found lesions in the olfactory organ as manifestations of copper poisoning. In Menidia menidia, dilation of blood vessels was apparent and haemorrhage had occurred in the brain and in periorbital connective tissue. Similar results are reported by Eisler & Gardner (1973). O’Hara (1971) found that copper concentrations in the medium had a strong correlation with respiratory response in juvenile bluegill. Above a concentration of 2.4 mg·1−1 the response changed and became more erratic. It was suggested that studies of this type could be used to determine permissible copper levels for an area. Hidaka (1970) examined the effect of a variety of transition metals on chemoreceptors of the carp. Copper, as CuSO4, tended to reduce the response of the receptor to both sugar and salt at the same time. Similar results were found for zinc, silver, and uranium. Platinum, gold, mercury, and telurium depressed the response to sugar. It was suggested that the mechanism of the action of the metal was a result of the interaction of the metal with a— SH group on the receptor site. Burton, Jones & Cairns (1971) showed that at least part of the toxic action of zinc was due to coagulation of mucus causing the suffocation of fish exposed to acutely toxic concentrations. Based on the results of Baker (1969), the phenomenon demonstrated by Burton et al. (1971) for zinc might be applicable to copper. Baker, however, found a decrease in mucous cells and an increase in chloride cells in winter flounder exposed to copper suggesting that less, not more, mucus was produced. In a literature review by Ohlendorf, Risebrough & Vermeer (1978) on the exposure of marine birds to environmental pollutants toxic effects were not noted for copper although very little is known about the biological importance of copper to marine birds. Anderlini, Connors, Risebrough & Martin (1972) determined concentrations of heavy metals in

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some Antarctic and North American birds and comment that “the two essential elements investigated, Cu and Zn, showed no significant geographical or species variation among the four populations of petrels. Distribution of individual levels about the mean for each tissue and each population was narrower than for the non-essential elements indicative of metabolically regulated levels for Cu and Zn.” Vincente & Chabert (1978) found low concentrations of copper in the blubber of a dolphin aground on the Mediterranean coast. Harms, Drescher & Huschenbeth (1978) examined a number of marine mammals from German coastal waters and comment that “copper and zinc have vital functions in the metabolism of higher developed animals and there is evidence that they can regulate the metal levels in their organs (Bryan, 1976). This may be the reason why no elevated copper and zinc levels have been found with increasing age in any of the marine organisms investigated.”

UPTAKE AND ACCUMULATION OF COPPER BY ORGANISMS The amount of copper, and other metals taken up by an aquatic organism is termed the concentration factor (e.g., Goldberg, 1965) and is related to the amount present in the water. As Burrell & Schubel (1977) point out, “this term is a useful parameter, for example, for phytoplankton and seaweeds, but should probably not be applied to seagrasses, since metals are believed to be largely derived from the sediment environment”. Harrison (in Templeton, 1978) points out that the quantity of copper accumulated by marine organisms depends on the concentration and chemical speciation of copper in the environment as well as the duration and pathway of exposure, and the ability of the organisms to regulate its uptake. Phillips & Russo (1978) comment that there is still much to be learned about the bioaccumulation of metals in aquatic organisms. The uptake of a metal, as with a nutrient, is dependent on many things (e.g., Bryan, 1975). Two of these are the surface area of the organism and the metal concentration in the environment (Coleman, Coleman & Rice, 1971; Varenko & Chuiko, 1971). In plants, a “coefficient of biological absorption” (“CBA”) has been obtained for the ratio between the content of a particular trace element in the plant, the surface area of the plant, and the level in the environment (Varenko & Chuiko, 1971). Seasonal changes in the concentration of copper in the environment will affect the CBA as will seasonal changes in the concentration of copper in the plant and its surface area. The residence time of the metal within the plant and the difference in the uptake and loss of the metal is dependent upon the physiology and metabolism of the organism (Wolfe & Rice, 1972) as well as the surface area and concentration in the environment. Seasonal changes in environmental factors such as temperature, salinity, metal concentration, and pH are also known to affect the rate of uptake of the metal (Bender, Huggett & Sloan, 1972; Wolfe & Rice, 1972). Concentration factors of metals in organisms have been found to follow the IrvingWilliams stability order for transition metal complexes (e.g., Goldberg, 1965; Elias, 1973), or to conform to the Freundlich adsorption distribution. The comments from a number of studies (e.g., Bryan, 1976; Burrell & Schubel, 1977) suggest that metal species

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partitioning of the copper must be considered when attempting to explain the organism concentration factor in theoretical terms. Schlesinger (1979) remarks how important it is to understand the entire fate of a pollutant within the environment as a whole in order to ascertain all possible hazards caused by its release. Arnac (1976) points out that “biologic-dynamic” conditions affecting the concentrations of metals may vary from one metal to another. The interaction of the organism with the environment and the sediments becomes important in understanding both the uptake of copper and the partitioning of the metal in marine and estuarine environments (e.g., Pellenbarg, 1978). General reviews on the bioaccumulation of heavy metals in marine organisms include those of Leland, Luoma & Fielden (1979) and Phillips & Russo (1978). Diagrammatic representations of metal speciation factors and metal uptake and utilization are given in Figures 1 and 2. Horne (1969) provides several concentration factors for copper in marine organisms which give an indication of the ability of the organism to concentrate the metal against a gradient. These factors (“fish”=80, scallop and mussel=3000, nudibranch=4300, oyster=13 700) suggest that there is a “uniqueness” to the organism and possibly some relationship to the proximity of the metal source. Andreev & Markov (1971) show accumulations of copper in intestinal parasites of up to 30–40 times that in the intestine of the host. This suggests an active uptake and concentration although Chaisemartin (1977) suggests that the copper level in tissues of the freshwater pearl mussel, Margaritifera margaritifera, approximates those of the environment and Morris & Bale (1975) suggest passive uptake of copper, cadmium, and zinc (but not manganese) in Fucus vesiculosus. Aubert et al. (1975), studying a benthic food chain, showed that there were different concentration factors with different chemical species of metals while Aoyama, Inoue & Inoue (1978a,b), working with 137Cs, found a sufficiently consistent concentration factor through a food chain to allow development of a model to predict heavy metal accumulation. Eisler (1979a) discusses the various factors affecting accumulation in coastal and marine biota, finding that filter-feeding bivalve molluscs generally contained the highest levels of copper while vertebrates had the lowest. Anderlini (1974) found no correlation between metal concentration and size in the red abalone (Haliotis rufescens) on the California coast. One interesting unexplained association however, was that the concentrations of copper tended to decrease in individuals in a north-south direction, while silver tended to increase, primarily in levels in the digestive gland and foot. A similar situation was found by Wenzloff, Greig, Merrill & Ropes (1979) with the surf clam (Spisula solidissima) and the ocean quahog (Arctica islandica). Cossa et al. (1980) propose that metabolic changes associated with sexual maturity and season continually change the relative importance of the circulating and storage compartments of metals and may be responsible for some of the variations that have been observed. The high concentration of copper and iron found in the hepatopancreas of the cephalopod Sepia officinalis has been associated with storage in cytolysosome-like dense bodies

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Fig. 1.—Sources and fates of copper in the marine environment.

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Fig. 2.—Sources and fates of copper available to marine organisms: reaction of copper with metabolites (3a–f) may occur within the cell, on the cell surface, or in the environment.

(Schipp & Hevert, 1978) and ultimate use in the synthesis of haemocyanin. Zaba & Harris (1978) examined uptake and effects of a number of cations, including copper, in fish liver mitochondria. They suggest that the interactions of copper and potassium could play a part in copper toxicity and that trout liver mitochondria have a high affinity for Cu2+ and may be a target for copper toxicity in fish. Morisawa & Mohri (1972) found that copper in sea urchin spermatozoa was restricted to the middle, between the head and the tail. There is considerable debate on the amount of accumulation, the conditions under which accumulation can occur, and the importance in economically valuable species. Majori & Petronio (1973), for example, found accumulation in the edible mussel (Mytilus galloprovincialis) but pointed out that the amount of mussel that would provide a lethal dose for an adult human was ≈534 kg. Not only is this a substantial meal but before anyone died from too much copper they would have ingested several lethal doses of mercury (14.7 kg of mussels) and more than one lethal dose of cadmium (294kg). Metal accumulation is frequently associated with organic ligands. Talbot & Magee

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(1978) isolated low molecular weight cadmium-, copper-, and zinc-binding proteins from M.edulis in a cadmium polluted area near Melbourne, Australia. They suggest that the molecular weight and other properties ally them with metallothionein-like proteins and that the synthesis may have been in response to the excess cadmium. Noel-Lambot, Gerday & Disteche (1978b) found metallothioneins in the liver of Anguilla anguilla, principally in the form of zinc and copper derivatives, but when the eel was exposed to cadmium the level of metallothionein increased and, during chronic intoxication, most of the metal was bound to metallothionein. Noel-Lambot, Bouquegneau, Frankenne & Disteche (1978a) compared the distribution of cadmium, zinc, and copper in three molluscs (Patella vulgata, Littorina littorea, Purpura lapillus) and found that the rate of accumulation varied according to the species and that metallothioneins may be responsible for high metal loads in some organisms living in polluted sea water. Olafson, Kearns & Sim (1979) discuss the induction of metallothionein synthesis in the hepatopancreas of the crab Scylla serrata and (Olafson, Sim & Boto, 1979) determined the amino-acid composition and found that the molecular weight, U.V. absorption spectra, isolectric points as well as the amino-acid composition were similar to that of vertebrate metallothionein. Rainbow & Scott (1979) describe the metal-binding capabilities of two proteins from the midgut glands of the crab Carcinus maenas while Viarengo et al. (1980) found copper-binding proteins in the gills of the mussel Mytilus galloprovincialis after 24 to 48-h exposures to sublethal concentrations of copper. Roesijadi (1981) discusses the significance of low molecular weight, metallothionein-like proteins in marine invertebrates and summarizes the available literature. Metallothioneinlike agents appear to be widespread. Lerch (1980) working with the fungus Neurospora crassa found that it accumulated copper with the concomitant synthesis of a metallothionein-like protein which bound a total of 6 g-at. of copper per mw of 2200. Rainbow, Scott, Wiggins & Jackson (1980), however, found no evidence for the binding of copper, zinc, and cadmium to metallothionein-like proteins in the barnacle Semibalanus balanoides suggesting that the mechanism is not common to all organisms. The concentration of copper in an organism is the result of metal uptake and storage. Nicholls, Curl & Bowen (1959) note variability in the levels of metals obtained through spectrographic analysis of marine plankton. They point out that there is considerable variation in metal levels given in the literature and that there are differences between individuals in different environmental conditions. It is also known that differences in technique provide differences in metal levels (e.g., Schmidt, 1978a,b) and that seasonal events can cause changes in metal levels in organisms (e.g., Fletcher & King, 1978a). These, as well as differences in metal forms from one geochemical environment to another and taxonomic problems with closely related species, all contribute to the variability found in metal levels reported in the literature. With the species specific nature of accumulation, misidentification of the organism with the resultant analysis of a mixture of species may be one of the many sources of variability. The concentration of some trace elements in the shells of molluscs has been associated with certain environmental factors (e.g., Rucker & Valentine, 1961). Paleoecologists have attempted to use this relationship as an index of conditions occurring in the past (e.g., Pilkey & Goodell, 1964) although recognizing certain problems. The effects of weathering or diagenesis may, as pointed out by Windom & Smith (1972), change the

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metal concentrations within the shells. Natural variations in the metals have also been suggested to be sufficiently great that environmentally influenced changes in concentrations within the shells will not be evident (Windom & Smith, 1972). In contrast, Kopfler & Mayer (1968) suggested a copper-zinc relationship in oysters which would suggest an order in the uptake. This is supported by Bender et al. (1972) who have shown a copper-zinc relationship in oyster shells from uncontaminated rivers in Virginia while oysters collected from areas of suspected unnatural inputs of metal do not show this relationship. Windley (1977) points out that algal fossils from some of the earliest geological records (some 3×109 years ago) contain copper, iron, nickel, and calcium. This, combined with the indication that algae such as Ascophyllum nodosum tend to accumulate metals as a function of environmental concentration and chemical state (e.g., Myklestad, Eide & Melsom, 1977) may allow some use of fossil materials to indicate metal conditions in the geological past. Even with the recognized variability of metal levels in organisms there does appear to be a relationship between metal concentration in the organism and in the environment. Stokes et al. (1973) found that laboratory and field strains of freshwater algae belonging to Chlorella and Scenedesmus exhibited a linear uptake with respect to the concentration in the growth medium. In contrast, Eyres & Pugh-Thomas (1978) found that copper concentrations in the tissue of the freshwater leach (Erpobdell octoculata) had an inverse relationship with substrate copper concentrations in a high pH stream. Ayling (1974) notes that, in Crassostrea gigas, “copper and chromium are accumulated by physiological processes that are governed by the size of the oyster and are almost independent of the concentration of metal at each site”. Greig, Nelson & Nelson (1975) examined the transfer of metals from adult oysters to eggs and found that the concentration of copper and cadmium in the eggs approximated that in adults. Boyden (1977) examined the metal content of several molluscs and found that the results could be affected by size of organism, the time required for the organism to come into equilibrium with the environment levels, and the nature of the environment. Bryan & Hummerstone (1978) examined heavy metals in the burrowing bivalve Scrobicularia plana from an uncontaminated and a contaminated estuary and found that copper increased with size in both estuaries although this was not the case for all metals. Marine molluscs frequently show a copper-size relationship although Foster & Bates (1978) found that the concentration of copper was inversely related to body weight in freshwater mussels used to monitor point source industrial discharges. Cross, Hardy, Jones & Barker (1973) found that the concentrations of metals in white muscle of the bluefish (Pomatomus saltatrix) and a bathyldemersal fish (Antimora rostrata) differed in their relation to size of organism. Concentrations of mercury increased with size while concentrations of manganese, iron, copper, and zinc either remained constant or decreased. They suggested that decreases found in the case of iron in the bluefish and zinc and copper in Antimora could be due to compositional changes in the muscle or changes in the diet. Yamamoto, Ishii, Sato & Ikeda (1977b) found that copper content in the hepatopancreas of carp decreased when ascorbic acid was added to the diet suggesting that a complexing agent could cause a decrease in metal accumulation. Fletcher & King (1978a) found a seasonal change in metals in gonads and liver of Pseudopleuronectes americanus but suggested that it was due not only to feeding

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and storage in the liver but, at least with copper, calcium, and magnesium, absorption of these metals from sea water during the non-feeding part of the year. There was also a difference in the accumulation of copper between testes and ovaries, the ovaries accumulating four to six times more than the testes which suggested the possibility of hormonal regulation of metal accumulation. Hormones have been found to indirectly affect the accumulation of copper in Ross’ Goose, Anser rossii; estrogen presumably accounts for the higher levels of copper in the primary feathers of the adult female as compared with those of the adult male (Hanson & Jones, 1974). Net uptake of copper in Mytilus edulis was found to be affected by seasonal variation, salinity, temperature, and the presence of other metals, as well as changes in the relative concentrations of the metals (Phillips, 1976a,b). Lunz (1972) found that the uptake of copper in Crassostrea virginica appeared to be a function of the species of the metal in the water which flowed over the mantle and gills. Lunz also noted a large variation in the ability to accumulate copper in oysters from the same population. Young oysters in an estuarine region in Australia were found to accumulate metal in relation to the geochemical nature of the estuary (Thornton, Watling & Darracott, 1975). Copper flux into the tissues of C.gigas and C.virginica appears to be dependent upon the copper concentration in the water however, suggesting little ability to metabolically regulate uptake (Chesapeake Research Consortium, 1972; Harrison & Rice, 1979). Kamimura (1980), however, suggests that the accumulation of copper and other heavy metals in oysters is primarily through food. In reviewing the literature on metal uptake it becomes apparent not only that there is environmental control on availability but also that there is a species specific response. Eide et al. (1979) monitored heavy metal pollution with in situ phytoplankton cage cultures in two Norwegian fjords and found that the growth and metal uptake of three species varied both between species and from place to place. In a field bioassay test for detecting contaminant uptake from dredge material by marsh plants Wolf, Gallagher & Pennington (1978) found that the nature of the dredge spoil tended to produce differences in metal uptake including copper. Ernst & Van der Werff (1978) found that copper uptake in Elodea nuttallii was dependent upon a number of environmental factors including copper concentration while Ko vacs (1978) andMudroch & Capobianco (1978, 1979) found that metal accumulation in freshwater aquatic and marsh plants was controlled by geochemical factors and the nature of the organism. Guthrie et al. (1979) noted that bio-magnification of heavy metals by organisms associated with a clump of oysters was species specific while Greig (1979) found differences in accumulation of copper in three species of molluscs, at each of three different levels of copper. Since interspecific differences exist in levels of accumulation, prediction of levels of accumulation requires information not only on the biological availability of copper but also on the nature of the mechanisms responsible for metal uptake. This information is almost non-existent in marine organisms although Seip (1979) developed a mathematical model for the uptake of heavy metals (primarily zinc) in Ascophyllum nodosum in which age-dependent growth factors were assigned. Black & Mitchell (1952), in reporting their work with brown algae,

indicate that concentration of trace elements may be due to ion exchange in which the concentration should vary with stage of development,

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normally a seasonal factor. Lunde (1970) indicates that the seasonal change in the amount of organic material in the algae is an important factor when combined with the change in concentration of the trace elements in the water. Surprisingly, although ample reason exists for the seasonal variation of copper in algae, the evidence available suggests that it is slight. Black & Mitchell (1952) found that copper, unlike some other metals, did not exhibit a seasonal trend although the techniques used caused them to denote the copper estimates as “semi-quantitative”. Young & Langille (1958) did not find a distinct seasonal cycle in copper values in Chondrus crispus, levels ranging from 14–35 µg·g−1 (dry wt) and ratios of concentration (sea water: plant) from 50–600 in a variety of green, red, and brown algae. Wort (1955) did not find noticeable seasonal fluctuations of copper in either Macrocystis integrifolia or Nereocystis luetkeana. The possibility that variation does occur may be considered, however, in the light of Lunde’s statement (1970) that seasonal “…variation is dependent on the form in which the trace element is present in the algae…”. Thus the partitioning of the metal within the plant may be important; concentrations in various forms may fluctuate seasonally while the overall concentration remains relatively constant. The difference in concentration of copper and other metals in benthic animals is suggested by Phelps, Santiago, Luciano & Brizarry (1969) to be dependent upon the metal-organism relationships, certain metals being concentrated in response to levels in food, others being concentrated independently from the food. Montgomery et al. (1978) found that the sea urchin Lytechinus variegatus and the holothurian Holothuria mexicana, feeding in an environment affected by sewage sludge, had a net uptake of copper while the clam Codokia orbicularis, the oyster Crassostrea rhizophora, and the snail Nerita tessplata did not. The concentration of metals in marine organisms has been considered to be partially due to the proximity of the organisms to a metal source. Establier (1969a,b) found a relationship between the copper levels in an oyster and the proximity of the metal source. Eisler et al. (1977a) found that clams (Pitar morrhuana) collected near a Rhode Island (U.S.A.) electroplating plant, characteristically had higher levels of silver, cadmium, cobalt, chromium, copper, iron, manganese, nickel, lead, and zinc compared with similarly-sized individuals collected further away from the plant. Foster & Bates (1978) found that caged freshwater mussels close to an industrial outfall accumulated more copper than those downstream of the outfall. Montgomery & Price (1979) found that metals leached from sewage sludge by flowing sea water, in a model turtle grassmangrove system, were taken up by organisms in the system. Leaves of the turtle grass (Thalassia testudinum) and roots of the red mangrove Rhizophora mangle showed a net uptake of metals, including copper, as did the sea urchin (Lytechinus variegatus) which grazes on Thalassia leaves. In a freshwater situation Cherry, Guthrie, Sherberger & Larrick (1979) found copper accumulation in sediments and most organisms when they examined the influence of coal ash and thermal discharges on bioaccumulation in aquatic

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invertebrates. Foster (1976) showed that variation in heavy metal content of Fucus vesiculosus and Ascophyllum nodosum was a function of the proximity of the organism to the source of the metal. Ireland & Wootton (1977) noted differences in the concentration of lead, zinc, copper, and manganese in the marine gastropods Thais lapillus and Littorina littorea and discussed them in terms of pollution sources. Huggett, Bender & Sloane (1973) suggested that the concentration of copper, cadmium, and zinc in Crassostrea virginica was related not only to the source but also to the position of the organism in the estuary with respect to water circulation which would control the distribution of the metal. Arnac (1976) found highest values of copper in water on the south shore of the St Lawrence Estuary; this was attributed to the circulation pattern. Bryan (1973) attributed the seasonal change in trace metal concentration in two species of scallops (Pecten maximus, Chlamys opercularis) to a change in activity and to the effect of increased phytoplankton productivity in the spring. Based on the high levels of metal in the digestive gland and kidney, tissues most likely to accumulate metals from food, Bryan suggested that metal concentration in scallops is associated with phytoplankton concentration. The change in activity is in part believed to be a function of temperature, controlling the rate of metabolic activity, and also of food availability. Although the effect of increased phytoplankton production is not apparent in the change in total copper concentration (Morris, 1971; Knauer & Martin, 1973), it is in terms of organically bound metal concentrations (Foster & Morris, 1971). Godoy & Barth (1967) found a relationship between copper concentration and biomass, increased productivity producing an increase in metabolite concentration in the water which increases the possibility of organic complexation of metals (Lucas, 1947; Provasoli, 1963; Barber & Ryther, 1969; Steemann Nielsen & Wium-Andersen, 1970; Lewis, Ramnarine & Evans, 1971). Seasonal variation in metal concentration in the environment and in organisms, especially sessile animals and plants, may also be related to the seasonal changes in land drainage and thus metal input into estuarine systems (Establier, 1969a,b; Bryan, 1973; Bender et al., 1972). There also tends to be a decrease in metal concentration in certain organisms as one proceeds away from a river mouth (Bender et al., 1972; Ireland, 1973; Peden, Crothers, Waterfall & Beasley, 1973). Petkevitch & Stepanyuk (1970) show that the seasonal variation of some trace metals in shrimp is very pronounced, especially during the breeding season, although variations in copper are slight and both intraspecific and interspecific differences in concentrations were noted. Djangmah (1970) showed that copper in the blood of Crangon vulgaris was closely related to the moulting cycle as well as to the season. Additional individual variation may be due to the nutritional status of the animals. Djangmah noted that during starvation, the copper concentration within the organism decreased and would rise again on a return to feeding. Macrobenthic organisms in intertidal and subtidal marshes have been found to be important reservoirs for copper and zinc (Kendall, 1978) suggesting metal uptake from both sediment and the water column. Kendall found that subtidal polychaetes accounted for >30% of the total copper budget. Bryan (1976) noted a near linear relationship between copper levels in the environment and Nereis diversicolor and suggests that absorption by most animals may involve passive diffusion of the metal as well as uptake from food. He indicates that metal uptake from food depends on the chemical form of the

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metal, that very stable metal species may not be broken down by digestion. Marsh sediments frequently contain elevated levels of copper (e.g., Gardner, 1976). Harris et al. (1979) examined the distribution of heavy metals in Western Port, Australia, and found the highest concentrations in sediments and seagrasses of a marsh apparently derived from an old swamp. Even within a marsh environment, however, concentrations of metals in sediments vary, Mudroch & Capobianco (1978) for example, found a high correlation between metal levels and organic carbon in sediments of marshes on Lake St Clair, Ontario, Canada. They also found variation in metal uptake by the same species of plants growing in different plant communities suggesting a control of metal availability within a marsh environment. Marsh grass plant communities form an important place where metal speciation is affected, not only as a result of metal uptake by the plants, but also the release of metals and organics through plant metabolism and degradation (Elias, 1973; Mathis, 1973; Smith, 1974; Windom, 1975; Pulich, Barnes & Parker, 1976; Wolf, Gallagher & Pennington, 1978; Burrell & Schubel, 1977; Pellenbarg, 1978). Schmidt, Wildung & Garland (1978) found five “classes” of copper-containing organic compounds in interstitial water from estuarine sediments, ranging from ionic species to materials with a molecular weight of >100 000. The major portion of the copper was associated with organics having a molecular weight of <500. Brooks & Rumsby (1965) found that the distribution of copper and silver in whole animals showed definite trends; where absorption of particulates is suspected copper levels are generally higher. This suggests that without any internal controls an organism intimately exposed to both particulate and dissolved copper will tend to accumulate higher levels than an organism exposed only to the dissolved fraction. J.S.Young et al. (1979b) found a threshold concentration for copper accumulation between 3–6 µg·1−1 in Eudistylia vancouveri. The body burden of copper increased above natural levels in areas of industrial discharge where copper levels were above the threshold limit. Although the suggestion is that any addition of copper beyond the threshold limit will cause an increase in body burden it is indicated by others (e.g., Sunda & Guillard, 1976) that it is the level of the biologically available species of copper that is important. Dodge & Theis (1979) for example, showed that uptake of copper by the freshwater larvae of Chironomus tentans did not occur when the copper-glycine and copper-NTA species were dominant but did occur when the free cupric ion and copper hydroxy complexes were present. From studies with dead specimens they conclude that the uptake is largely passive involving chemical interactions between Cu2+ and absorption sites at the surface or interior of the organism. Naturally occurring agents such as humic materials or Gelbstoff (Milanovich et al., 1975) will affect the level of these species and perhaps the threshold should be in terms of biologically available species of copper, not total or dissolved copper. Seasonal changes in run-off (e.g., Sankaranarayanan et al., 1978) will also affect the total levels of metals as well as the levels of the various chemical species. Pellenbarg (1978) found that the surface microlayer in a salt marsh was enriched by organometallic compounds formed from decomposition of the marsh grass Spartina alterniflora. In any environment in which decomposition is taking place this needs to be considered; Patrick & Loutit (1977) noted that removal of bacterial epiphytes from a freshwater plant in a polluted river reduced the concentrations of metals significantly. G.E.Jones et al. (1976) found that two marine bacteria, Arthrobacter marinus and

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Pseudomonas cuprodurans, assimilated copper and other metals against a gradient. These studies suggest that bacterial activity may provide a mechanism for trace metal speciation and biological uptake. The ability to accumulate metals exhibited by aquatic organisms has been utilized in the removal of heavy metals from sewage treatment plant effluent. Filip, Peters, Adams & Middlebrooks (1979) found that semi-continuous cultures of the mixed algal flora native to wastewater lagoons absorbed 70–90% of the cadmium and copper. Windom (1977) tested the ability of salt marsh vegetation to remove nutrients and heavy metals from dredged material disposal area effluents. Copper was found to accumulate more in sediments than in Spartina alterniflora roots or leaves suggesting that the 15–32% metal removal efficiencies obtained were due more to the effect of the sediments than the vegetation. Vestergaard (1979) has found a relationship between the metal content of salt marsh soils and the content in the water. Greig & Wenzloff (1978) found that the copper level in oysters maintained in sand-filtered water did not increase whereas those maintained in unfiltered water did suggesting the importance of sedimentary material in controlling the biological availability of copper. The amount of copper taken from the water into biological material is not only a function of the nature of the environment and the physiological requirements of the organism, but also the amount of biological material. In the Menai Strait (U.K.) Morris (1971) noted changes in zinc, copper, and manganese which he associated with a phytoplankton bloom (Phaeocystis). Uptake of metals by organisms, however, means that a greater amount of the metal is in the particulate state and less is in the dissolved state. With the death and decay of the organisms the metal is returned to the water, although sinking of the particle will cause vertical movement to the sediments. Wolfe et al. (1976) found that a significant amount of one year’s accumulation of iron, manganese, copper, and zinc in sediments in an estuary was that which had been incorporated into eelgrass (Zostera marina). Martin (1970) found that with the high concentration of copepod crustaceans in the oceans and with the number of moults that the copepod has during its lifetime, these organisms and their moults will be an important factor in the biological transport of trace metals. Osterberg, Pattullo & Pearcy (1964) commented that the “… biological transport of zinc across the pycnocline…[by migrating euphausiids] is undoubtedly more important than transport through this layer by physical processes”. Small & Fowler (1973) estimated the time to cycle zinc in the Ligurian Sea, through the existing euphausiid populations, as between 498 and 1243 yr and calculated the net vertical flux of zinc by migrating euphausiids. This movement is estimated to account for 81–98% of the total transport of zinc. Uptake of copper tends to occur more readily at certain sites or in certain regions and tends to be from metal in certain geochemical configurations. In planktonic autotrophs and some types of heterotrophs, uptake is directly from the “dissolved” fraction in the water column while in plants such as the mangrove, uptake is from both the water column and interstitial water (Banus, 1975). Metal uptake in organisms which ingest particulate organic material can occur either through the food or across the body surface. Harrison (in Templeton, 1978) states that primary sites of copper absorption and accumulation seem to be membranes in respiratory and digestive organs. White & Thomas (1912) suggest that fish take up dissolved copper most readily through the gills and J.S.Young et

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al. (1979b) suggest that the branchial crown of sabellid polychaetes forms an important site for copper uptake. In other organisms, such as oysters, uptake can occur through the entire body surface (Harrison & Rice, 1979; Chesapeake Research Consortium, 1972) although differences in uptake may be found between various body surfaces. Saenko et al. (1976) found selective accumulation in the liver, gills, gonads, and Keber’s organ in molluscs and in liver, stomach and peritoneal epithelium in the starfish. Metal uptake may be passive, as in the case of Fucus vesiculosus (Morris & Bale, 1975) or active as in the case of the marine bacterium Arthrobacter marinus (G.E.Jones et al., 1976). Myklestad et al. (1978) examined the changes in levels of zinc, lead, cadmium, and mercury in Ascophyllum nodosum transplanted from an area with high to an area with low metal concentrations. They found that the metal content of the plant mirrored the change; older plant material grown in the original high metal area retained most of its high levels, young plant material grown in the low metal area had low levels of metals. Khovrychev (1973) examined absorption of copper ions by cells of the yeast Candida utilis and found that the rate of absorption could be described by the MichaelisMenten equation but that absorption of copper was inhibited in the presence of mercury and lead salts. Nielsen (1976) found that copper uptake by barley could also be described by the Michaelis-Menten equation. The form of the metal that is actually taken across the cell membrane is discussed for terrestrial plants by Tiffin (1977) who comments that metals may be taken up either passively or actively, the latter either by pH modification or release of organic agents. If the latter the trace metals bound to organic agents in transport were mostly in the anionic form and this suggested chelation by organic and amino acids. Brooks & Rumsby (1965) give three possible pathways for the concentration of trace elements by bivalves: (1) ingestion of inorganic suspended material (e.g., Armstrong & Atkins, 1950); (2) ingestion of elements after pre-concentration in food material (e.g., Bowen & Sutton, 1951); and (3) complexing of metals by coordinate linkages with appropriate organic molecules (Schubert, 1954), the incorporation of metal ions into physiologically important systems (Lehninger, 1950; Williams, 1953), and uptake by exchange e.g. onto mucous sheets of the oyster (Korringa, 1952). Goldberg (1957) notes that the order of enrichment of divalent metal ions by members of the marine biosphere followed the order of stability of metal-ligand complexes. This was supported by the work of Cross, Duke & Willis (1970) who examined the distribution of manganese, iron, and zinc in sediments, water, and polychaete worms. It was contradicted by the work of Morris (1971) as well as by Bowen (1966) on the basis of the affinities exhibited for metals by a variety of organisms. Bowen points out that the heavier elements in the group IIA and group IA metals tend to be more readily taken up than the lighter elements. Harriss (1965) found that the Ni:Co and Ca:Sr ratios in algae were distinctly different from the same ratios in sea water while the K: Rb ratio was only slightly different. This led Harriss to suspect that metabolic processes were responsible for selectivity of chemically similar metals. Jenne & Luoma (1977) suggest that it is the thermodynamic activity of the uncomplexed ion that is probably the most important factor controlling the biological availability of solute trace elements. There is some evidence to indicate, however, that under conditions of metal stress (Zn, Cd, Cu) the permeability of the cell membrane changes (Yager & Harry, 1964). This would reduce

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the ability of the organism to control metal balance under stress. Bryan (1976) suggests that absorption of metals from solution involves passive diffusion with permeability being important in determining metal tolerance. He also suggests that the availability of metals from food depends on the chemical form of the metal, and that accumulation may be a result of the metal species in the food. Canterford, Buchanan & Ducker (1978) studied the accumulation of heavy metals by the marine diatom Ditylum brightwelli and found that, with copper, there was no definite correlation between metal concentration and metal taken up. Delisle, Hummel & Wheeland (1975) found that the rate and degree of accumulation of heavy metals in organisms exposed to sediments contaminated with such metals was dependent on such factors as the chemical form and concentration of the metals, the particle size and agitation of the sediment, and the feeding habits and physiological characteristics of the organisms. Luoma & Bryan (1979) comment, with respect to the biological availability of zinc, that availability to deposit-feeders is determined by the partitioning of the metal among substrata. Harrison & Rice (1979), however, found that copper flux (as 64Cu) into tissues increased with copper concentrations in the water. In an examination of the availability of sediment-adsorbed heavy metals to benthos Neff, Foster & Slowey (1978) measured the accumulation of cadmium, chromium, copper, iron, manganese, nickel, lead, and zinc by five species of test organisms. Statistically significant accumulation of metals was found only 36 times out of 136 metal-species-sediment test combinations and significant accumulation of copper occurred only in one species (Neanthes arenaceodentata), in one sediment type, at one salinity. Variations in bioaccumulation of metals were observed between species, metals, sediments, and salinity; bulk metal analysis of the test sediments did not correlate with metal accumulation. The uptake of metal ions by marine organisms may be via a mucous membrane. Korringa (1952) found that a number of positive polyvalent ions like Al3 +, Cu2+, Fe2 +, Zn2 +, Hg2 +, and Mn2+ were easily caught and accumulated by the mucous net of the oyster, but not the positive monovalent ions like Na and K although present in greater quantities or the negatively charged ions. In the case of the oyster the mucus is ingested, exposing it to digestion and transport processes in the digestive tract. The ingested metals may not be totally absorbed, selection being possible as a result of membrane response or chemical nature of the metal (e.g., Harrison 1977a,b). Selectivity is also suggested by elevated concentrations of metals in faecal material (Boothe & Knauer, 1972). Davies (1972) states that “the presence of abnormally high levels of an element such as zinc within the diet may reduce the absorption of an element such as copper, if the level of copper is marginal with respect to the nutritional requirement”. The metal complex of the food and the nature of the intestinal tract has been shown to affect both directly and indirectly the availability of metals for absorption in terrestrial (Davies, 1972) and marine (Bryan, 1976) animals. This may control the nature and concentration of metals taken into the blood stream but has less effect on incorporation of metals into the cell. Davies (1972) feels that competition for absorption of the metal is minimal in comparison with competition for incorporation into the cell. The ability of some organisms to control metal levels in the tissues (Olafson & Thompson, 1974) supports this concept although the apparently “open” nature of other organisms (see Bryan, 1976) suggests that the concept does not apply to all animals. The presence of a metallothionein-like complexing

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agent in some marine organisms (Olafson & Thompson, 1974) provides a mechanism for internal control of metal levels. The actual mechanisms involved in transferring copper and other metals from the environment into the organism, to and into the cell, and to the site of utilization within the cell form an extremely important link in organism-metal relationships. Understanding the nature of the mechanisms that, for example, may allow regulation of manganese but not cadmium, copper, and zinc in the alga Fucus vesiculosus (Morris & Bale, 1975) is critical to an explanation of the effect of copper on Fucus. Schuster & Broada (1970) found that the cell wall of Chlorella was capable of binding zinc reversibly; two types of active sites were found, each with its own affinities and capabilities. Wassermann (1949) examined ion exchange properties of alginates from marine algae and found different selectivity coefficients for cadmium-strontium ion exchange reactions. The response of alginates to divalent ions appears to be dependent on the nature of the ion (Haug, 1961) and follows the ionotropic series of Thiele: . Neilands (1974) reviewed the iron transport compounds found in micro-organisms and identified a series of iron-containing metabolites or siderochromes which “…fulfill a unique and essential function in regulating the iron nutrition of the microbial cell”. Information of this type does not exist for copper-containing compounds. The effect of copper can be modified by the addition of complexing agents (Jones, 1970; Lewis et al., 1972, 1973). At least with some phytoplankton, the complexing agent controls the level of cupric ion which has been found to be an important species of copper (e.g., Sunda & Guillard, 1976). Although the action of complexing agents occurs primarily outside the cell, the control of the concentration of various copper species in the medium exerts control over the amount of copper available for biological uptake. Organically bound copper has been found in marine environments (Corcoran & Alexander, 1964; Alexander & Corcoran, 1967; Slowey, Jeffrey & Hood, 1967) and metabolites such as amino acids are suspected to be of importance (Siegal & Degens, 1966) especially because the levels of organically bound copper increase during periods of high productivity (Foster & Morris, 1971). (See pp. 539–552 for a discussion of complexing agents.) The problems of determining the nature of the active agents include the low concentration in sea water as well as the presence of a variety of organic compounds. Only recently has some headway been made in the isolation and identification of complexing agents in natural waters. Mantoura, Dickson & Riley (1978) found that, in general, the order of increasing strength of binding of metals by isolated humics followed the Irving–Williams series. The specificity of the metal-organic reaction forms an important phase of study (e.g., Duursma, 1970; Siegal, 1971). Yager & Harry (1966), for example, found that the synthetic chelating agent ethylene diaminetetraacetic acid (EDTA) inhibited the uptake of zinc by the freshwater snail Taphius glabratus but did not inhibit uptake of cadmium. Owens & Chaney (1971) found that Chlorella reduced weakly chelated Fe3+ to Fe2+ before uptake into the cell but was capable of taking up Fe2+ without change, apparently in both ionic and chelated forms. The addition of a strong chelator, with Fe3 +, prevented the cells from reducing the ion to the required Fe2 +. Once within an organism, copper has three possible fates: (1) to participate in some metabolic process; (2) to be returned to the environment; and (3) to form associations that

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store or immobilize the metal thus reducing its effect on the organism. Participation in metabolic processes is discussed on pp. 412–494. Measurement of the return to, or exchange with the environment has been difficult to study, partly because of the short half-life of the radioisotopes of copper. As a result, the cycling of copper through biological material has been difficult to measure. Liberation of copper and other metals from plankton and benthic forms has been suggested by a number of workers (e.g., Grave, 1941; Glaser & Anslow, 1949; Pyefinch & Downing, 1949; Whitfield & Lewis, 1976). Seeliger & Edwards (1979) found that the copper in two benthic red marine algae was found in the thalli and associated with organics and dissolved organics released from the living thalli. Peter, Welsh & Denny (1979), working on translocation of lead and copper in two submerged freshwater angiosperm species, found extensive acropetal translocation of copper to accumulation sites in the stem, especially in the stem apices, and in the youngest leaves. Danielsson (1980) found a strong association of copper with the silica portion of diatoms suggesting that the metal may not be incorporated into the cell itself. Martin (1979) found high levels of copper in the “liver” of the squid, Loligo opalescens, suggesting some concentrating mechanism. He also found elevated levels of silver and cadmium in the same organ which he states “…may cause copper to be concentrated in the liver; this process thus reflects a sub-lethal effect caused by these heavy metals”. It is also quite possible that a biochemical mechanism such as metallothionein causes the concentration of a number of metals in the hepatopancreas. Howard & Nickless (1978) found that copper and cadmium in Littorina littorea were in association with a low molecular weight extract which supports the hypothesis that bioaccumulation within organisms is due to one or more organic complexing agents. Shuster & Pringle (1968) note a loss of copper from Crassostrea virginica and suggest a direct loss to the environment. They found that at the end of an experiment on the accumulation of copper by the oysters “…most of the oysters ‘bled’ globs of blue green material (presumed to contain copper) resembling semi-coagulated vertebrate blood”. The same authors commented on the observation of Galtsoff (1964) that a large proportion of the copper in oysters was in blood cells and suggested that it may be in a proteinaceous complex. Coombs & George (1977) also found vesicles in both Mytilus edulis and Ostrea edulis which contained heavy metals, including copper, formed under conditions of excess metals, as a mechanism for immobilizing the metals. Ikuta (1968) found that oysters with high levels of copper and zinc lost the excess metal after 116 days in a normal environment. Scrudata & Estes (1976) found that the concentration of copper and zinc in Crassostrea virginica tissues was dependent on the weight of body fluid. A change in body fluid could account for at least part of the loss that they observed in tissue metal levels. Doyle, Blake, Woo & Yevich (1978) found phosphorite concretions containing high concentrations of metals (Cu=4000 µg·g−1) in the kidneys of two molluscs, Mercenaria mercenaria and Agropecten irradians. They comment that these appear to be a normal formation of the excretory process of molluscs under certain kinds of stress and that, through geological time they may change into crystalline structures similar to marine phosphorite deposits which otherwise cannot be explained by the chemical precipitation-replacement hypothesis. Rheinberger, Hoffman & Yevich (1979) propose that elemental analysis of the concretions found in Mercenaria mercenaria as

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well as other histopathological features of the molluscan kidney may be a convenient method of tagging the various heavy metal pollutants in the water-sediment environment to which the benthic community is exposed. If the copper in oysters is in blood cells, as Galtsoff (1964) as well as Ruddell & Rains (1975) indicate, this forms either a mechanism for transport or a means of immobilizing the metal. Ruddell & Rains suggest that the oyster removes copper from the environment only as required, which would mean that if it were immobilized it would be in ‘anticipation’ of future needs. Martoja, Tue & Elkaïm (1980) found copper, in the form of CuS needles, inside the lysosomes of Littorina littorea. Accumulation of copper sulphide increased with the age of the animal, was independent of environmental levels, and was interpreted as being the result of haemocyanin degradation. Walker (1977) found copper granules in the parenchyma cells of the prosoma of Balanus balanoides and suggests that the immobilization of copper by barnacles is an example of an internal detoxification process for heavy metals. Olafson & Thompson (1974) found metallothionein-like proteins in several marine organisms. These compounds form a mechanism for immobilizing heavy metals and allowing some excretion. Joyner & Eisler (1961) noted a tendency for salmon fingerlings to concentrate radioactive zinc in the vertebral column, head, and visceral mass. They found that the outer surface of the bone appeared to serve as an ion-exchange medium capable of taking up large quantities of metal ions and that “metals thus exchanged from serum proteins to the bone may be prevented from undergoing further exchange by the overlayering action of the growing bone”.

SOURCES OF COPPER AND CHANGES OF POTENTIAL BIOLOGICAL IMPORTANCE OCCURRING IN COPPER SPECIES IN ESTUARINE AND MARINE ENVIRONMENTS The forms or species of copper entering the marine environment are not necessarily those encountered by members of marine food chains. Changes which occur in the form of the metal are due to changes in the nature of the medium as well as the effect of organisms within the medium (e.g., Carpenter, Bradford & Grant, 1975; Fig. 1). It is the sources of copper and these changes that are discussed in this section as well as the importance that the changes in speciation have on controlling the biological availability of copper in estuarine and marine environments. Nriagu (1979b) discusses the copper contents of the various reservoirs of the world, the transfer of copper between the reservoirs, and the effect of human activity on the system. Major sources of copper are from the atmosphere (Duce et al., 1974; Peirson, Cawse & Cambray, 1974; Peyton, McIntosh & Anderson, 1974; Duce, Ray, Hoffman & Walsh, 1976; Boyle, Sclater & Edmond, 1977; Cattell & Scott, 1978), geothermal sources, and run-off from terrestrial sources (e.g., Spencer & Sachs, 1969; Turekian, 1971). Jedwab (1979) has described metallic brass and copper from the open ocean indicating the technical influence of particulate copper in the oceans and Roskam 1972) describes a case of acute fish toxicity on the coast of Holland, from several kilograms of copper sulphate crystals. The oceans also serve as a source for atmospheric copper in aerosols.

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Ericksson (1959, 1960) and Blanchard (1963) estimate an annual input of 1015 to 1016 g of copper into the atmosphere. Boyle et al. (1977) suggest that aeolian input is comparable in magnitude to fluvial input and is important in nearshore environments and Lantzy & MacKenzie (1979) estimate a continental dust flux of copper of ≈1010 g·yr−1 with a combined industrial particulate emission plus fossil fuel mix of 2630×108 g·yr−1. Cambray, Jefferies & Topping (1979) estimated that the Rhine introduces 2000 tonnes·yr−1 of copper into the North Sea while rain water introduces 5600 tonnes·yr−1 copper. The flux of copper is not just from the atmosphere into the oceans but, through bursting bubbles and aerosol transport, from the oceans into the atmosphere. Piotrowicz, Duce, Fasching & Weisel (1979) suggest that the sea may contribute 10% or more of the total annual quantity of copper to the atmosphere. Kazaikin, Mel’Nik & Podgorskaya (1978) present data on the distribution of heavy metal pollutants in the atmosphere close to the surface of the water and Kretzschmar & Cosemans (1979) discuss the levels of heavy metals in the air at the Belgian North Sea coast during May 1972–April 1977. East to southeast airflows, from heavily industrialized regions, give the highest levels for all heavy metals while north to northwest winds, from the North Sea, bring the lowest levels. Examination of the windrose for copper shows input from both European mainland and England. Airborne copper originates from a variety of sources although smelters are frequently considered as the major ones (e.g., Buchanan, 1973; Beckman, 1978). Nriagu (1979a) discusses the chemical and physical properties of copper aerosols and the removal of copper from the atmosphere by various mechanisms. Although atmospheric flux may not control deep ocean particulate chemistry (BuatMenard & Chesselet, 1979) it is important to the surface microlayer (e.g., Slinn et al., 1978), which contains material from the atmosphere as well as from the water. Once in this layer copper is available for vertical movement into subsurface waters as well as aerosol input into the atmosphere. Brass (1980) notes that the celestite (SrSO4) skeletons of acantharians may be an important transport vector for trace metals, extracting them from surface water and releasing them at the depth of the oxygen minimum-nutrient maximum. Copper may be a lattice cation in the celestite and has an estimated removal rate from the surface water of 40 ng·cm−2·yr−1. In an examination of the trace element geochemistry of the continental shelf waters of the southeastern United States, Windom (1978) estimated that sedimentation from the water column caused the accumulation of 6mg of copper·m−2·yr−1 on the bottom sediments of the continental shelf, that 3 mg·yr−1 is removed from the dissolved phase by biological mechanisms from a one metre square water column of average depth and that an estimated 1.5 mg of copper is introduced each year by atmospheric mechanisms to every metre square of sea surface. Pellenbarg & Church (1979) calculated trace metal budgets in a marsh near Lewes, Delaware, U.S.A. and found values of copper in the microlayer to be ≈10% of the total metal flux. They noted that organic materials from Spartina litter form an integral part of the microlayer and that such organic materials may provide natural chelators for trace metals. Wallace, Hoffman & Duce (1977) examined the effect of organic matter and the importance of atmospheric deposition on the particulate trace metal concentration the surface water of the northwest Atlantic and suggest that atmospheric input of particulate copper, although important, is not the sole source. They estimate atmospheric input to the mixed layer as 10 ng·cm−2·yr−1 while that from the mixed layer, via particulate organic material, was 33

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ng·cm−2·yr−1. Metal-rich surface microlayers have also been noted in fresh water (Elzerman & Armstrong, 1979) and Elzerman, Armstrong & Andren (1979) suggest that a significant part of the metal may be atmospherically transported and anthropogenic. The mechanisms involved in the transfer of atmospheric trace materials past the air-sea interface have been excellently reviewed by Slinn et al. (1978). Fluvial input tends to be more localized than aeolian input and may also be affected by metals in sewage (Galloway, 1972b; Goldberg, 1972; Greene, 1976; Gross, 1976; Reish, Martin, Piltz & Word, 1976; Bertine & Goldberg, 1977; Hershelman, Jan & Schafer, 1977; Higgs, 1977; Schafer, 1977; Summerhayes et al., 1977; Thorell, 1977; Bascom, 1978; Duedall et al., 1978; Griggs & Johnson, 1978; Soulsby, Lowthio & Houston, 1978; Simps on, 1979) with proved or suspected effect on metal concentration in organisms (Greene, 1976; McDermott et al., 1976; Reish et al., 1976). Galloway (1979) calculated the global copper input from municipal waste-water outfalls to be ≈42000 metric tons·yr−1 while that from natural weathering is 250000 metric tons·yr−1. Galloway (1979) points out that, in southern California, metals in the waste water are primarily associated with the solid fraction although <10% of the metals are retained in contaminated sediments in the outfall area. The remainder are either dissolved out of the solids suspended in the water or are transported out of the area while still associated with the particulate fraction. Metal levels in seafood organisms near southern California outfalls have been measured by Jan, Moore & Young (1977) with copper values suggesting that there is little concentration in invertebrates but some in fish. Montgomery & Price (1979) also found an accumulation of copper in the sea urchin Lytechinus variegatus when it fed on the leaves of Thalassia testudinum exposed to metals leached from sewage sludge in a flow-through model system. Amiel & Navrot (1978) examined the effect of trace metals in sewage on metal levels in nearshore sediments while McGreer (1979a, b) found a relationship between the concentrations of mercury and cadmium, but not copper, in sediments near a sewage outfall and the burrowing response of Macoma balthica. Bertine & Goldberg (1977), using metal/aluminium ratios in sediments in the basin between San Clemente and Santa Catalina Islands, some 70 km off the coast of southern California, found no evidence of introduction of anthropogenic copper although there was evidence of input of several other metals and copper is introduced along the coast. Oliver & Cosgrove (1974) found that metal input into a sewage treatment plant was not a continuous process but occurred as “slugs” lasting for a discrete period and Lantz (1979) examined a heavy metals waste treatment plant which used co-precipitation with hydroxide. Lantz commented that, especially when sea water is present, the discharge criteria could not be met by the process but found that the use of soluble sulphide and a pH of 7.5 allowed removal of metals tested (except hexavalent chromium) to the required effluent goals. Ackermann & Waitz (1976) discuss the use of ion exchange resins to reduce heavy metal concentrations of waste water. The sludge fraction of sewage produces problems not only because of its high metal content but also because of the cost of treatment. The United States of America alone produces five million dry tons of municipal sludge each year and this is expected to double by 1990 (Heritage, 1978). Ocean disposal of waste-water sludge accounts for ≈ 15% of the current sludge volume (Anonymous, 1978). This, however, may end in the U.S.A. by 1981 (Bartos, 1978) even although there is some belief that the effect of ocean

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dumping of sewage sludge has only a slight or no detrimental effect on the marine ecosystem (Guarino, Nelson & Almeida, 1979). Daillaire (1979) points out that the concern with sewage sludge is the presence of high levels of heavy metals but he comments that bioaccumulation has not been proved in many areas. Input of copper into estuarine and marine environments is also affected by metals in industrial effluents and urban run-off (Gross et al., 1971; Skei, Price & Calvert, 1973; Armstrong, Hanson & Gaudette, 1976; Sommer, Diachenko, Pyzik & Siegrist, 1976; Goldberg, Gamble, Griffin & Koide, 1977; Knauer, 1977; Koch, Hall & Yesaki, 1977; Lichtfuss & Breummer, 1977; Vivian & Massie, 1977; Christensen & Scherfig, 1978; Dominik, Förstner, Mangini & Reineck, 1978; Goldberg et al., 1978a; Griggs & Johnson, 1978; Randall, Grizzard & Hoehn, 1978; So, 1978; Williams, Baldwin & Robertson, 1978). Loring (1978c) found that concentrations of copper in bottom samples from the Saguenay Fjord in the Gulf of St Lawrence, Canada, decreased seaward and that sources were both natural and industrial. McGrath & Austin (1979) found elevated levels of zinc and to a lesser extent copper in Belfast Lough, when compared with a non-industrialized area of the Irish Sea. It is, however, frequently difficult to differentiate between the biological effects of copper introduced in rivers and copper originating in cities close to the mouth of the rivers but introduced directly into the estuary (e.g., Bryan & Hummerstone, 1977). Hung, Li & Wu (1975) found increased levels of copper in Kaohsiung Harbor (Taiwan). Gardner, Chen & Settlemyre (1976) compared polluted and pristine marsh sediments in South Carolina and found higher levels of copper in the polluted sediment but commented that “…natural processes, as opposed to industrial contamination, can adequately account for these results”. This was partially based on the similarity of the Cu:Zn ratios in the two types of sediments and the indication that either salinity or organic productivity could be a reason for the greater level in the contaminated sediment. Taylor (1979) examined the effect of discharges from three industrialized estuaries on the distribution of heavy metals in the coastal sediments of the North Sea and found that of the metals studied, copper, manganese, and zinc were found at increased levels in the area studied compared with a control area. Near the mouth of the Tees Estuary, the geology of the area appears to be a more important factor than the industrial input in determining the metal content of the marine sediments. Vestergaard (1979) suggested that salt marshes may indicate the degree of trace metal pollution of protected coastal waters and noted that “the amounts of Pb, Cu, Zn and Ni extractable with EDTA are a measure of the amount of the non-lattice fractions present in the soil”. The suggestion that appreciable levels of copper are introduced by all industrial operations should not be drawn from this review, rather that only certain types of industry form sources of the metal. Jackson(1979a,b), for example, found that copper was not present in concentrations significantly above background levels near an actively worked oil field in the northwestern Gulf of Mexico and was not accumulated in either the sediments or the organisms in the region. This has also been found by Newbury (1979) and Wheeler, Schwarzer & Anderson (1978). Anderson et al. (in Wolfe, 1978) examined the ‘bioavailability’ of hydrocarbons and heavy metals to marine detritivores from oilimpacted sediments and found that the crude oil used (Prudhoe Bay Crude oil) had <5 µg copper·g crude−1 and that, compared with the normal levels of trace metals in marine sediments, the uptake by the deposit-feeders tested (Phascalosoma agassizii and Macoma

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inquinata) would be negligible. Bulk loading of copper ore has been found to produce increased levels in sediments near the loading facility; Chapman, McGreer & Vigers (1979) found sediment levels of up to 14% copper beneath a bulk loading area in Vancouver, Canada, although they found that there was “a healthy and diverse community of organisms living on and around the pilings”. Furthermore, the results of a 96-h LC50 bioassay with coho smolts and “shore crabs” suggested that the copper was not in a biologically available state. Goulder et al. (1979a) found that high metal concentrations (including copper) and low pH in water from the Humber Estuary, U.K. adjacent to a metal refinery plant reduced bacterial activities. Eisler et al. (1977b) found increased levels of several metals, including copper, in sediments from Narragansett Bay during the autumn of 1973, presumably from an electroplating facility on Quonset Point that operated from 1943–1973. They also found high levels of copper in Pitar morrhuana near the plant. Olthof (1978) discusses the fate of heavy metals from metal finishing plants and land disposal of solid wastes and suggests that disposal to the ocean floor appears to be the least hazardous method for the environment, for while aesthetically not pleasing, neither plant nor animal life is affected, and there is no measurable recycling of the metals. It is also suggested that “…the storm water leachate from a sludge deposit (on land), after the mother liquor has been washed out, essentially has the same metal content as any natural body of water, river, or lake”. There are also biologically important localized inputs of copper due to the use of copper tubing and sea water as a coolant in power plants or as a source of water in desalination plants (Zeitoun, Mandelli & McIlhenny, 1969; Chesher, 1971; Mandelli & McIlhenny, 1971; Romeril & Davis, 1976; Harrison, 1977a,b; Martin et al., 1977). Compton & Corcoran (1976) examined the levels of copper in various parts of the circulating systems of 26 different power plants with freshwater cooling systems and found higher levels in re-circulating waters, blowdown or spill-over waters and discharge to settling ponds but an appreciable reduction in copper during pond residency. They suggest that properly designed cooling towers would produce no significant pollution of natural water bodies from copper. Trace metal speciation has also been examined in saline waters affected by geothermal power plants (Sposito & Page, 1977) which introduce biologically important levels of metals (Tullis, 1977). As an example of the effects of corrosion of cooling coils, approximately one ton of copper is lost yearly from the copper tubing in a sea water desalination plant in the Gulf of Eilat, Israel (Le-ver, Gat & Friedland, 1976). Young, Jan & Moore (1977) examined metal levels in cooling water discharges of a number of power plants in California and found an average difference between effluent and influent copper concentrations of 0.21 µg·1−1 for dissolved copper and 0.10 µg·1−1 for particulate copper. Although there was a net addition of copper they point out that none of the increases in any of the metals examined exceeded 0.2 µg·1−1 (sic) and that part of the additional metals in the thermal effluent may have been contributed by organisms swept into the intake systems of the plant. If the whole body levels of copper in these organisms were of the order of 10 mg·g−1 (a typical value), and the concentration in such organisms in the influent is 10 µg·kh−1 extraction of the copper from organisms during the cooling process would raise the level of dissolved copper in the coolant sea water by 10−10. It should be noted, however, that retention basin waters were not discharged at the time the samples were

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being taken for the study which would remove one major source of metals. From studies of fallout from plants remote from marine areas and using fresh water for cooling purposes, it is also apparent that fossil fuel plants release copper from the fuel into the atmosphere with subsequent fallout or by leaching from ash disposal facilities (Theis & Richter, 1979; Theis, Westrick, Hsu & Marley, 1978; Klein & Russell, 1973; Wiener, 1979) as well as from cooling pipes (Delfino, 1977). The biological effect of any copper introduced by power plant cooling water discharges is affected not only by the increase in temperature of the cooling water during its passage through the system but also by the use of chlorine as an antifouling agent. Mattice (1977) notes that the toxicity of chlorine is affected by temperature, pH, copper, and nickel. Hoss et al. (1974) found a synergistic effect of copper and temperature on larval, postlarval, and juvenile pinfish (Lagodon rhomboides), flounder (Paralichthys sp.), and mullet (Mugil cephalus) although the level of copper used (1 mg·1−1) was unreasonably high. Harrison (1977a, b, c, 1979a) and Harrison et al. (1977), in a series of quarterly reports on the biological effect of copper from nuclear and fossil fuel power plants with salt water cooling systems, noted an increase in copper, especially in particulates. Harrison et al. (1977) found that some organisms were affected by copper concentrations as low as 30 µg·1−1. Because copper concentrations in power plant effluents have frequently exceeded toxic levels, the use of surface water as a receptor for wastes may have already resulted in adverse effects on nearshore populations. It must be realized, however, that levels as high as 30 µg·1−1 are not common, that they are usually the result of cleaning with chlorine, and that the organism would have to live within the heat exchanger pipe to experience that level of copper and then only for short periods. What must be considered under these abnormally high levels is that the effluent is diluted once it enters the receiving waters, with water having unique physical and chemical properties. Harrison (1979a) found that the receiving water had a copper complexing capacity approximating or exceeding the total copper in the effluent. Thorhaug & Schroeder (1978) found that temperature affects cation uptake in Thalassia testudinum in a predictable fashion in the 20–30 °C range although copper was not one of the metals tested in the evaluation of the effects of substances emitted from power plants in tropical and subtropical regions. Environmental degradation near an oil-fired steam generating plant in Puerto Rico could be related initially to temperature and chemicals including the accumulation of trace metals (Lopez, 1968). One other important source of copper is antifouling coverings (Groover, Lennox & Peterson, 1970; Hochman, 1972; Young, Heesen, McDermott & Smokler, 1974; Gerchakov & Sallman, 1977; Bellinger & Benham, 1978). Ikemura (1969) examined corrosion of copper alloys and found that when there is biofouling corrosion increases due to the production of localized anoxic zones as a result of organism decomposition with the resultant formation of corrosive hydrogen sulphide or ammonia. Syrett, Wing & MacDonald (1979) found that the presence of sulphide or one of several sulphide oxidation products turns non-corrosive de-aerated sea water into a comparatively corrosive environment causing the release of copper from copper-nickel alloys. Release of copper can also occur as a result of flow structure, through the erosion-corrosion process on copper-nickel alloys (Leumer, Schack, Graham & Perkins, 1978).

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Copper has long been incorporated into paints to reduce fouling on marine vessels, the idea being to provide a long term leaching of enough copper to produce a toxic environment on or next to the painted surface. D.R.Young et al. (1979) examined vesselrelated contamination by copper and other metals of southern California harbours and calculated an application of more than 180 metric tons per year of copper as copper bottom paints in southern California alone. They also found that Mytilus edulis near vessel scraping and re-painting yards contained up to ten times the natural levels of copper. Young & Alexander (1977) also found a difference in Mytilus edulis from different harbours which they suggested might be due to flushing characteristics. For the discussion on biological effects of copper from antifouling materials see pp. 560–562. Copper bound in marine sediments could also be considered a ‘source’ if remobilization occurs through dredging (DeCoursey & Vernberg, 1975; NOAA, 1975; Trefry & Presley, 1976; Morton, 1977). The very nature of the operation, however, means that, under most conditions, it is the exposure of former sediment-associated metal in a localized area and not the introduction of metal new to estuarine and marine environments. Studies in both fresh and salt water (e.g., Keillor & Ragotzkie, 1976) indicate that copper values may be high in the dredged material but that the metal is not bioaccumulated, even with sediment elutriates (Lee, Lopez & Mariani, 1975). Slowey & Neff (in Herbich, 1977) found that only those metals that are soluble in the interstitial water and readily released to the water column appear to be available to sediment-water interface feeders such as “Rangia cuneate” (sic) and Palaeomonetes sp. Studies with true deposit-feeding infauna, such as the Tubifex and Neanthes worms, indicate that there is little if any uptake from sediments from several areas when the organisms’ guts were purged prior to analysis. Slowey & Neff also comment that it is so far not possible to arrive at an extraction procedure that will predict the availability of heavy metals to the benthic-feeding infauna. This has also been noted by Hirsch, DiSalvo & Peddicord (1978) in a study on the effects of dredging and disposal on aquatic organisms. Windom (1977) indicated that most of the metal in effluent from dredged material is accumulated by sediments rather than by organisms and that inorganic chemical and physical processes probably account for much of the removal. A last source of biologically important copper is the input of metal from mining (e.g., Thompson & McComas, 1974; Thompson & Paton, 1978; Portmann, 1971; Padan, 1971). Discussion of the biological effects of copper from both dredging and mining is given on pages 562–567. THE FATE OF COPPER IN ESTUARIES Numerous papers have appeared concerning the changes that occur in the form of copper as it enters the marine environment from fluvial sources (e.g., Schmidt, 1978a, b). Turekian (1978), with reference primarily to radioisotopes, discusses the fate of metals in estuaries and the sources and processes affecting trace metals as they enter an estuary, particularly in the region of the salt-water-freshwater interface. The many sources of copper as well as the chemical differences between and within each type of source, make it impossible to draw broad generalizations on changes occurring during and after the introduction of copper. In addition, the receiving waters change with time and space

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which further complicates the explanation of copper speciation as it does with the response of organisms to copper. The biological effect of introduced copper is thus controlled by a large number of variables (Bawden, Heath & Nanton, 1973; Benedict, Hall & Koch, 1973; Baudo & Varini, 1976; Davey, 1976b; Cross & Sunda, 1977; Estuarine Pollution Control and Assessment, 1977; Ragsdale & Thorhaug, 1980) and specific effects range from metal accumulation (e.g., Stenner & Nickless, 1975; Thornton et al., 1975) to modifying the adaptations of anadromous fishes to salt water (Lorz & McPherson, 1976, 1977). The knowledge that inshore waters can accumulate high levels of both metal and hydrocarbon has been at least partly responsible for the siting and design of large industrial complexes (Amsden, Sweetin & Treilhard, 1978) and large research facilities to examine the biological effects of pollutants (Menzel & Case, 1977; Davis, Hester, Yoakum & Domey, 1977). Gibbs (1973) examined the transport mechanisms of several trace metals (Fe, Ni, Cu, Cr, Co, Mn) in two major rivers, the Amazon and the Yukon, both without any appreciable industrial pollution. The majority of copper was found in crystalline sediments (Amazon=74.3%, Yukon=87.3%) with lesser amounts being adsorbed to particulates (Amazon=4.9%, Yukon=2.3%). It was also found as metallic coatings on particulates (Amazon=8.1%, Yukon=3.8%), in particulates of biological origin (Amazon=7.6%, Yukon=3.3%) or in solution (Amazon=6.9%, Yukon=3.3%). The large percentage of copper found in crystalline sediments underlines the importance of the particulate phase of run-off entering estuaries. Sankaranarayanan & Reddy (1973) found that the copper content of a river was inversely proportional to the volume of flow suggesting a constant level of introduced copper. The nature of the source varies, however, from river to river and even within a drainage basin such that, under some conditions, run-off and copper content may be positively correlated (e.g., Girvan, Tatro & Hodgson, 1978). Martin & Meybeck (1979) derived an elemental mass-balance of material carried by the major rivers of the world and suggest that copper enriched deepsea clays “represent an admixture in various proportions of river-borne eroded material and marine biomass, with an additional supply of volcanogenic and atmospheric material”. Payne & Pickering (1975) found that kaolinite clay suspensions removed the cupric ion from solution although the extent of removal varied with solution pH, the nature of any ligands present, and the order of contact of the species. Wagemann, Brunskill & Graham, 1977) examined the reactivity of some river sediments from the Mackenzie Valley (Northwest Territories, Canada) with Beaufort Sea water (salinity=22.3‰). Both iron and manganese were released from the sediments to the sea water to a limited extent while copper, lead, and zinc were adsorbed from the sea water by the sediments. Stern (1975) found rapid adsorption of copper and lead by sediments (<62.5 µm) used to simulate dredge spoils. Farrah & Pickering (1978) found that the affinity for metals varied between clays while McDuffie, El-Barbary, Hollod & Tiberio (1976) found trace metals to be concentrated on smaller sized particles in a river system. Van der Weijden, Arnoldus & Meurs (1977) discuss desorption of metals from suspended material in the Rhine Estuary and state that both suspended matter and filtered water from a tidal station at Hoek van Holland are depleted of most heavy metals. They comment that in this area suspended matter and adsorbed and dissolved heavy metals do not show a conservative

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mixing behaviour. Both particulates and dissolved organic compounds have the ability to control metal speciation in fresh water and it is important to realize that rivers transport not only the metal but also the ligands. Pagenkopf (1978), for example, discusses the effect of humics and fulvics on metal-ion transport in fresh water while Hopner & Orliczek (1978) suggest that precipitation of river borne humic material in estuaries forms an important contribution to organic matter in the sediments and, through complexation, that it may exert important biological control over heavy metals. Wilson (1978) discusses an equilibrium model which describes the influence of humic materials on the speciation of Cu2+, Zn2+, and Mn2+ in fresh water. Sholkovitz, Boyle & Price (1978) found that dissolved humic acids were removed from river water at quite low salinities and (Sholkovitz, 1978) that this was associated with the flocculation of dissolved iron, manganese, aluminium, copper, nickel, cobalt, and cadmium. Surface films formed from allochthonous surfactants may concentrate trace metals in foams in fresh water (Pojasek & Zajicek, 1978). Bergman, Ritter, Zamierowski & Cothern (1979) found that particular organic compounds in sewage, as defined by density, contain the greatest concentration of trace elements, including copper. Duinker & Kramer (1977) examined the speciation of dissolved zinc, cadmium, lead, and copper in River Rhine and North Sea water and found dissolved organic material to be associated with the apparent complexing ability of the water. (Inorganic complexation, colloid formation, and adsorption may have also been partially responsible for the results that were observed.) The nature of the river with respect to the type and concentration of both organic compounds and particulates dictates the nature of the metal species reaching the estuary. Bergmann, Lehnen & Seehaus (1978) estimated transport and partial metal loads for the River Ems Estuary, taking into account some of the tidal cycle effects. Brzezinska (1979) examined the run-off of cadmium, copper, lead, and zinc in the waters of the River Vistula from 1st January, 1976 to 1st March, 1977, in order to estimate the annual discharge into the Gulf of Gdansk, Baltic Sea. These budgets are useful in providing input data although perturbations from local or atmospheric sources (e.g., Huggett & Bender, 1977) may provide important short term changes in the amount of input as well as the nature of the metal. Copper introduced into estuarine and marine regions or re-introduced via dredging undergoes some change due to variations in the chemical nature of the environment. Some of the effects of the processes controlling metal speciation are discussed in a UNESCO publication (1979). Leckie & James (1974) as well as Suffet (1977) discuss the nature of aqueous metal ions, their interaction with solids, and transport processes in natural aquatic systems. Burton (1978b) provides a general review of metal speciation associated with the sedimentary cycle in estuaries. Hosokawa, Ohshima & Kondo (1970) suggest that when river water mixes with sea water in an estuary, the concentration of the dissolved elements may be changed by one of two general mechanisms. The first is a simple mixing process while the second is described as a series of complex chemical reactions. Montgomery & Santiago (1978) note that the highest concentration of organically bound dissolved copper is in the mixing zone of the estuary. Turekian (1977) suggests that the humic-metal colloids which give the appearance of a dissolved state in fresh water, flocculate when mixed with sea water. Abdullah, Royle & Morris (1972)

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comment that the amount of metal entering an estuary combined with mixing and circulation control the level of that metal. Riley (1937) suggests that mixing and resultant precipitation as well as dilution are of primary importance in determining the concentration of copper in the Mississippi Delta. In the transition zone between river and estuarine water there is precipitation and flocculation of both organic and inorganic particles with associated adsorption and desorption of metals (e.g., Cossa & Poulet, 1978). Demina, Gordeyev & Fomina (1978) suggest that the principle association of copper is with the inorganic compounds, and that it is predominantly in the “dissolved form”. In a study of mixing, removal, and mobilization of trace metals in the Rhine Estuary Duinker & Nolting (1978) suggest that the concentrations of trace metals, including copper, in the bottom sediments of the Estuary are due to mixing and sedimentation processes in the water column. They suggest that the Estuary retains trace metals, and that only a small part of the dissolved and particulate metals entering the Estuary through the Rhine River reach the ocean. Eaton (1979b) examined the geochemistry of soluble copper, iron, nickel, and zinc in the San Francisco Bay Estuary and concluded that physical mixing processes for the most part govern variations of soluble trace metals at average to slightly below average run-off discharges. Eaton also found that maximum copper removal at low run-off, occurs in the region of the turbidity maximum. Morris, Mantoura, Bale & Howland (1978), however, suggest that the apparent decrease in dissolved copper may be due to a change in speciation, into an analytically unavailable form such as copper-organic complexes rather than simply to sedimentation processes. This is also suggested by Burton (1978a,b) and is of biological importance because the effect of the metal will be different in an organic complex than as an adsorbed ion (Schlesinger, 1979). In an examination of the distributions of dissolved and particulate trace metals in the estuary and Gulf of St Lawrence, Hoff (1978) found that the concentrations of Fe, Mn, Cu, and Zn were increased both by St Lawrence River water and, at least with copper and zinc, by upwelled subsurface water. Loring (1978a,b) noted that copper was a potential contaminant of the upper St Lawrence Estuary. Hunter (1980) found surface microlayer enrichment in the North Sea adjacent to an area receiving fluvial input of clay minerals and attributed this enrichment either to the vertical movement of metals from flotation of particles attached to rising bubbles or to atmospheric input, or to both processes. The suggestion from a series of studies is that salinity and flow are important in controlling both the concentration and state of the metal. Dissolved metals in river water, especially those in ionic form, may precipitate upon entrance into salt water (Lowman et al., 1966). Sundararaj & Krishnamurthy (1972) and Taylor (1976) found a negative correlation between salinity and copper concentration in near-surface estuarine waters. Batley & Gardner (1978) found that metal distributions were consistent with other measured physical and chemical variables in an Australian estuary. Cutshall, Holton & Small (1973) found that copper, zinc, and manganese were not conserved during mixing of Columbia River water with sea water, that metal concentrations at ≈7‰ salinity were higher than expected, whereas precipitation of metal occurred at ≈20‰ salinity. Girvin et al. (1978) found that the partitioning of copper between the dissolved and particulate phases was affected by river discharge and salinity.

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With the radioisotope 65Zn, exchange reactions have been shown to occur with particulates (Johnson, Cutshall & Osterberg, 1967; Aston & Duursma, 1973). Turekian (1971) noted a general loss of metals from particulates as they entered the estuarine environment. Thomas & Grill (1977) noted increases in the concentration of dissolved metal in the Fraser River Estuary (Canada) which could not be attributed to the level of dissolved metals in the river and were apparently due to desorptive exchange reactions when the river water mixed with sea water. Grieve & Fletcher (1976, 1977), however, noted that there is adsorption of metals to hydrous oxides in the seawater-freshwater mixing zone of a river and commented that this may be an important estuarine process which allows accumulation of metals in nearshore sediments. In addition, Burrell (1977) and Heggie & Burrell (1978) found that copper was removed from soluton by particulates and transported to the bottom in an Alaskan fjord estuary. Evans (1978) suggested that copper concentrations are controlled by adsorption-desorption reactions with manganese oxides. De Groot (1970, 1973, 1976) noted that adsorption of metals to particulates occurred in estuarine situations but that the formation of metal-organic complexes mobilized some of the metals and lead to less contaminated sediments in the lower courses of the river deltas. Bopp (1977) suggested that only 1–5% of all metals associated with sediments are in the adsorbed phase in the Delaware River although he noted that the adsorbed phase is environmentally active. It is apparent that there is a need for more information on the adsorption-desorption reactions of copper in estuaries and the biological availability of the resulting metal species. Salomons & De Groot (1978) found that copper was most abundant in the acidoxidizable fraction of sediments in the Rhine River Estuary. Serne (1977) found that, of the copper in San Francisco Bay sediments, 53% existed in mineral lattice sites that are essentially inert to moderate chemical attack, 43% was associated with organic and sulphide-like phases and a small percentage appeared to be associated with hydrous oxides, probably by adsorption. Menon, Ghuman & Emeh (1980) examined trace element release from estuarine sediments in a lagoon near the J.F.Kennedy Space Center in Florida, finding that 59% of the copper was associated with the crystalline structure while 30% was deposited as part of the metallic coatings around the particles. It is important to relate the various metal fractions to conditions in estuaries to determine the fate of the metal. Bewers & Yeats (1979), for example, found that dissolution of trace metals adsorbed onto iron and manganese oxides will accompany reduction of these oxides in the sediments. The formation of relatively insoluble sulphides may be expected in reducing sediments, but these trace metal sulphides are sufficiently soluble to give concentrations of pore water much higher than those of the overlying sea water. The metals may then cross the sediment-water interface by diffusion, thus causing metal enrichments of bottom waters. The exchangeability of the cations between the dissolved and particulate states is dependent upon the specificity of the adsorption to the particles and the solubility of the cation, insoluble forms remaining bound to the particles (Johnson et al., 1967). The rate of reaction is dependent not only upon the degree of change in environmental conditions but upon the properties of the metal itself. Copper has been found to precipitate in an almost linear fashion with time after its introduction into sea water, with <20% of the metal precipitated after 3 h exposure under laboratory conditions (Lowman et al., 1966).

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It should be remembered, however, that salinity can be a controlling factor as Cutshall et al. (1973) noted that there was no precipitation below a salinity of ≈20‰. The rate of precipitation of metals will exert some control over the concentration of the metal in the sediment. Iron, which precipitates rapidly, was found to decrease in concentration in the sediments proceeding offshore while copper was found to precipitate much less rapidly and levels tended to remain much more constant in the sediments (Lowman et al., 1966; Cross et al., 1970). Bradford (1976) found that the concentrations of a number of metals in the benthic sediments near an outfall in San Francisco Bay were controlled by metal-specific processes. Sieburth (1971) suggested a loss of humic materials in estuaries and indicated that trace metals complexed by the humic material would also be precipitated. Dehlinger et al. (1974) examined the fates of heavy metal wastes in parts of Long Island Sound and found that certain parts of the Sound act as sinks for sediments and metals and that the disposition of the metals was controlled by estuarine circulation. Olsen et al. (1978) noted that tidal and estuarine processes cause the accumulation of recent, polluted sediment in specific areas in the Hudson Estuary. Duinker & Nolting (1976) suggested that the Rhine Estuary acts as a sink for particulate trace metals entering from the River Rhine or formed after the introduction of metals in a dissolved state. Landing, Feely & Massoth (1978) examined major and trace element composition and vertical fluxes of two size fractions of particulate matter collected in sediment traps in an inshore area in Alaska. Copper was found to be enriched in the fine fraction although, due to the greater large-particle flux, the coarse fraction accounted for slightly more than 65% of the particulate copper flux. The relationships between particulates, organics, and copper (e.g., Giesy & Briese, 1978) whether in small or large streams, affects the nature of the copper introduced into an estuarine environment. Paul & Pillai (1978) found that copper was preferentially concentrated in water over the sediments of a river in India but that tidal variations and seasonal, monsoonal effects change the concentration as well as the distribution of copper in the river. Curtis (1979) found seasonal occurrences of trace metals in the Miami River (Florida) which were attributed to differing hydrological conditions due to the “aquifersurface” water interaction. Bewers, Macaulay & Sundby (1974) and Yeats, Bewers & Wallon (1978) examined the trace metal concentrations in waters of the Gulf of St Lawrence and the Scotian Shelf and estimated the amount of anthropogenic input necessary to increase the levels in receiving waters. They did not, however, find evidence of significant modification of levels of iron, manganese, cobalt, nickel, copper, zinc, and cadmium in eastern Canadian coastal waters. They suggest that monitoring of river waters will provide a more rapid estimation of anthropogenic activity than the monitoring of coastal waters. McDuffie et al. (1978) found downstream copper values higher than upstream values in the Susquehanna River bottom sediments and, from the ratios of silt: sand and clay: sand copper values, indicate that the higher copper levels are associated with particles of smaller diameter. Jones & Jordan (1979) examined the distribution of organic material and trace metals in sediments from the River Liffey Estuary in Dublin and found that although the surface enrichment of copper and several other metals may be due to an industrial input “diagenetic changes within the sediments may be responsible for maintaining the high levels at the surface”. From a study of trace metal concentrations in run-off from a heavily populated area

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Galloway (1972a,b) estimated that the introduction of anthropogenic metal was of the same order of magnitude as injection from natural weathering. Even with this, the effect of dilution of dissolved metals and mineralization of sediment associated metals was found to maintain fairly low values over a large geographic area. Salomons & De Groot (1978) in an examination of the trace metal pollution history in sediments of the Rhine River, found an increase in copper delivered to the sediments during 1921–1975. Lichtfuss & Bruemmer (1977) studied the sediments from the Elbe River and its Estuary and noted that from Hamburg to the mouth of the river, the heavy metal content decreased rapidly as a result of mixing polluted fluvial sediments with scarcely polluted marine sediments. Part of the reason for the low levels of sediment associated metals noted by Galloway (1972a,b) and by Lichtfuss & Bruemmer (1977) could be the effect of organic complexing agents which could reduce the level of sediment-associated metal (e.g., De Groot, 1970). Zirino, Leiberman & Clavell (1978b) noted a tidal fluctuation in the level of copper in San Diego Bay suggesting some concentrating or diluting mechanism which produces fluctuations in metal levels. Sick, Johnson & Engel (1978) found that a tidal front tended to accumulate metals in Delaware Bay, with concentrations of selected trace metals in zooplankton, particulates, and the dissolved fraction significantly higher in samples collected at the front. They suggest that advective transport by convergent “water masses” may produce the concentrating mechanism. Vertical movement of particulate copper introduced into the marine environment by rivers is due primarily to sedimentation processes. Formation and vertical transport of particulate organic matter and trace metals in open-ocean waters occurs through biological processes as well as bubbles (Wallace & Duce, 1978a). Wallace & Duce (1978b) found an upward bubble flux of copper of 1.4×10−5 µg·m−2·sec−1. They also found formation of large aggregates due to the effect of bubbles, causing the formation of particulate metals with their subsequent downward transport. Piotrowicz et al. (1979) found copper enrichment of aerosols through bubble transport to the surface and the release of metal-containing solution with the bubbles bursting at the air-sea interface. Copper enrichment appears to be correlated with biological processes in near surface waters and is influenced by both microlayer and scavenging effects. Any change in the input of copper into estuarine environments through time should be shown in the sediments unless there is bioturbation or vertical movement and refluxing. Skei & Paus (1979) found that they could calculate the average anthropogenic flux of zinc, lead, and copper for a dated sediment core from Ranafjord (Norway). The input due to anthropogenic sources for copper in the upper 10cm of the core was 20 mg·m−2·yr−1 which corresponds to sediments deposited since 1900 when mining activities started in the area. Heggie (1978) and others, however, have found that copper can be remobilized from the solid phase(s) in surface sediments and subsequently returned to the overlying water which would make the estimates of Skei & Paus (1979) somewhat lower than the actual metal input may have been. Perturbation of the sediments due to abnormal events will also produce a change in the copper levels. The 1972 tropical storm Agnes, for example, caused an appreciable increase in the sediment copper levels of the normally saline section of the Patuxent River which empties into Chesapeake Bay (Chesapeake Research Consortium, Inc., 1973). It was suggested that the major source of new copper

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was from the Chesapeake Bay, by adsorption or other surface reactions on mineral grains which were exposed to fresh water during the storm, and then re-exposed to salt water. An analogous process has been described by Grieve & Fletcher (1977), for the low salinity region near the mouth of the Fraser River (Canada). Particulates in the salt wedge extending up the river accumulate iron and other metals, by adsorption, from metals brought down by the river and precipitated upon exposure to low salinity conditions. The adsorption of metals such as zinc, in the case of Grieve & Fletcher (1977), or copper occurs either jointly with iron or because of the new surface formed by the iron coating, as a result of continued suspension of the particulates due to estuarine circulation in the mouth of the River. Besides precipitation-co-precipitation mechanisms and adsorption and sedimentation processes, the concentration of metals by organisms ultimately leads to some sedimentation and removal from the water column (e.g., Turekian, 1971; Horvath, Harriss & Attraw, 1972; Morozov, Patin, Demina & Tikhomirova, 1976). Horvath et al. (1972) suggest that the concentration of metals by plants with subsequent release during decomposition, in an estuarine environment, may be of some importance in controlling water quality. This is also suggested by Pellenbarg (1978) and Windom (1975) with Spartina litter in a salt marsh. Smith (1974) discusses the rôle of sea grasses and benthic algae in the geochemistry of trace metals in Texas estuaries. Gardner (1976) examines the exchange of nutrients and trace metals between marsh sediments and estuarine waters, and Hallberg (1974) suggests that copper in the sediments is complexed by agents derived from organic matter. Mathis (1973) found that red mangrove leaf detritus contains higher levels of metals than living leaves and suggests that it is due to adsorption, complexation, and concentration of dissolved metals from the water. Gardner & Kitchens (1978), in a review of chemical exchanges between salt marshes and coastal waters, note that even although the salt marsh system is important in metal exchange, the amount of copper, zinc, and molybdenum retained by the marsh relative to the flux through the salt marsh is sufficiently small to prevent direct measurement with existing techniques. The concentration of metals in sediments is, at least in part, a result of post depositional reactions in the sediments (Duchart, Calvert & Price, 1973; Brooks, Presley & Kaplan, 1968). In examining the effect of pulp and paper mill effluent on the St Croix River and Estuary in New Brunswick (Maine) Fink, Pope, Harris & Schick (1976) found that sequestering of most of the anthropogenic metals in sediments occurred but they suggested that these metals may be released at a later date. Turekian (1971) suggested that the decay of organics in sediments under anaerobic conditions holds the metals as sulphides or reduced oxides although Hallberg (1974) found that heavy metals in the reduced, hydrogen sulphide-rich layers of the sediment were not only present as sulphides but were also adsorbed or absorbed by clay minerals and organic material. Strom & Biggs (1972) found no significant change of copper with the depth of cores taken from a marsh near Lewes, Delaware, suggesting some stability of the metal once it is in the sediments. Hallberg (1974) and Elderfield et al. (1975) found, however, movement of metals due to diffusion or percolation of water through the sediments. It has been assumed that heavy metals immobilized in the bottom sediments constitute a potential hazard to water quality and aquatic life since perturbation and chemical

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change may cause their release (Forstner, 1979). A number of studies (e.g., Neff et al., 1978; Windom, 1975) have, however, indicated that the copper in sediments is usually not in a biologically available state. In addition, the distribution of copper and zinc in the sediments does not necessarily indicate the distribution of metal in biological tissue; Huggett, Cross & Bender (1975) for example, found that the copper concentration gradient in oysters did not appear to be related to the distribution of copper and zinc in the sediments. Cross et al. (1970) found intraspecific differences in concentrations of manganese, iron, and zinc in polychaete worms suggesting either the ability to regulate trace metal uptake or the control of biological availability by properties of the sediments. The ability of some organisms to exist in estuarine sediments containing high concentrations of heavy metals (Bryan & Hummerstone, 1971; Anonymous, 1972) may be a combination of the ability of the organism to exist in the high levels of metal and the nature of the metal species that are present. The speciation of copper in estuarine and marine waters is dependent on a number of physical and chemical factors including both inorganics and organics (e.g., Dryssen & Wedborg, 1974; Schmidt & Gibson, 1978; Davis & Leckie, 1979). The interaction of dissolved organic material with metals is complex (e.g., Moore, Burton, Williams & Young, 1979; Shuman & Cromer, 1979) and, with estuarine regions forming important areas of biological production, the effects of specific trace metal species on succession of natural populations of marine phytoplankton need to be determined in order to better develop predictive capabilities of the effects of anthropogenic inputs into coastal waters (Feeley & Curl, 1979). The control of copper speciation, whether in the water column or in the sediments, relates directly to the control of the biological availability of the metal. The partitioning of metal through the interaction with particulates or dissolved organics affects the ability of the organism to accumulate copper. The presence of humic compounds and clay minerals in river run-off and the change that occurs upon entry into sea water thus become biologically important. The presence of naturally occurring complexing agents in sea water (e.g., Corcoran & Alexander, 1964; Siegel, 1971) and the interaction of these agents with those being introduced in run-off can produce a complex system. The unique nature of the organism which dictates its ability to control metal uptake (Marks, 1938; Cross et al., 1970) is another complicating factor. Uptake of metals by organisms is, however, intimately associated with environmental conditions (e.g., Establier, 1969). S.E.Jørgensen (1979) discusses models that describe the distribution and effect of heavy metals in an aquatic ecosystem, with equations that describe variation in concentration in given trophic levels, the exchange of toxicant between sediment and water, and the equilibrium between dissolved and suspended states. Kazarian (1978) discusses the possibility of predicting the spread of radioactive fallout, with systems that would be useful in predicting the distribution of other anthropogenic materials. Leckie & James (1974) discuss control mechanisms for trace metals in natural waters and state that “it is the view here that phenomenological models are useful if they allow prediction, but even more important didactically, models provide a scientific metaphor that allows us to frame questions necessary to gain greater insight into natural phenomena”.

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METAL SPECIES AND BIOLOGICAL EFFECTS Copper occurs in both dissolved and particulate phases in marine waters and sediments. Within each of these phases there are numerous chemical and geochemical forms or species (e.g., Goldberg, 1976; Drever, 1977; Burton, 1978a; Demayo, Davis & Forbes, 1978; Zingaro, 1979). This is a result of copper associating with dissolved or particulate organics, inorganics or combinations of organics and inorganics. Reviews of literature dealing with species of metals include those of Turekian (1971), Leland et al. (1975), Minear et al. (1975), Olofsson & Ghosh (1975, Stumm & Brauner (1975), Leland, Copenhaver & Wilkes (1975), Reish et al. (1977), Leland, Luoma, Elder & Wilkes (1978), Reish et al. (1978), Schmidt (1978a,b), and Pickering (1979). The report of the workshop on copper in estuarine, continental, and marine waters (Templeton, 1977) as well as that on trace metals and marine production processes (Feely & Curl, 1979) point out the importance of understanding the physical and chemical nature of the metal species in order to understand the biological effects of copper in marine and estuarine waters. Batley & Florence (1976b) developed a scheme of classification of heavy metal species in natural waters which allowed grouping of species for quantitative assay purposes, a scheme comparable with that developed for fresh water by Tessier, Campbell & Bisson (1979). Mericam & Astruc (1979) discuss the nature of inorganic copper species in both fresh and salt water. Sugimura, Suzuki & Miyake (1978) discuss the chemical forms of minor metallic elements in the ocean, using XAD-2 resin as a mechanism to isolate the organic forms. Hoover (1978) reviews the literature dealing with the ecological significance of inorganic species of metals in fresh and salt water and the analytical requirements for their study. The biological activity of the various metal species has been compared to their chemical activity (e.g., Goldberg, 1957). The term “availability”, when applied to metals, can imply the ability to extract the metal chemically (e.g., Frazer et al., 1978) or the ability of the metal to react with the organism (e.g., Sunda & Guillard, 1976). Nygaard & Hill (1979) compare methods for the determination of “available” trace metals in sea water and use the term “availability” to indicate that metal collected or measured by various techniques. In conclusion they suggest that it is dangerous to equate “available” metals, as defined by analytical procedure, with biological “availability” and, in general, the term should be used to indicate that level of metal reacting with a specific technique, agent, or organism. It must also be realized that there is species specificity with organisms and, within a given species of organism there may be age effects or effects associated with physiological events as well as adaptation (e.g., Murphy & Belastock, 1980). In general, the physicochemical factors that regulate solubility are important in determining bio-availability (Patrick, Gambrell & Khalid, 1977) as is the nature of the association between copper and dissolved ligands (e.g., Lewis & Whitfield, 1974). Recent studies have included modelling of the relationships of copper with both dissolved and particulate materials in sea water and fresh water (e.g., Wilson, 1978) in an attempt to estimate speciation in terms of biological effect. Application of both modelling and speciation studies is not only in the estimation of biological effects of metal pollution but, as Cross & Sunda (1977) indicate, in the establishment of water quality criteria.

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Jenne & Luoma (1977) suggest that the biological availability of trace elements is related to the thermodynamic activity of the element while Goldberg (1976) discusses heavy metal speciation and biological effects. From a study of the toxicity of copper to Daphnia magna, Magnuson et al. (1978) suggest that the carbonate copper complexes are not toxic, the anionic hydroxo-copper complexes contribute 15–18% to the toxicity of copper, and that free copper and/or the neutral and/or cationic hydroxo-copper complexes are responsible for 60–70% of the toxicity. Whitfield & Turner (1979), in an assessment of the relationship between the biological, thermodynamic, and electrochemical availability of lead, suggest that the rate processes associated with uptake, release, and assimilation are relevant in the determination of the biological availability of a metal. Tiffin (1977) suggests that plants may take up trace metals passively or actively, the latter by modifying pH or releasing organic agents; transport of metals within plants is as a complex and mostly in the anionic form suggesting chelation by organic and amino acids. This section deals with the major groups of copper species that have been examined in terms of biological effect. In most cases the nature of the species is only inferred; very little is known about the chemical nature, especially with respect to natural organometallic compounds. Although the details in this section partly overlap those of the other sections it is designed to examine the geochemical processes in terms of the biological effect of copper in marine and estuarine waters. DISSOLVED COPPER When one discusses “dissolved” or “particulate” copper the problem is to provide a good definition of the terms. With the adsorption of copper to colloids and the complexation of copper by macromolecules, any definition ends up being simply a working definition. For most purposes, however, passage through a 0.45 µm filter is used as “dissolved” although it must be realized that changes between the dissolved and particulate state occur frequently (e.g., Horvath, 1973). The formation of stable species may occur through adsorption by particulates (e.g., Aston & Duursma, 1973), complexation by inorganic ligands (e.g., Zirino & Healy, 1970; Zirono & Yamamoto, 1972), and complexation by organics (e.g., Barber & Ryther, 1969; Barsdate, 1970; Bender, Matson & Jordan, 1970). Changes in the species of copper are, however, produced by numerous factors (see Bezborodov & Zhorov, 1977). The rôle that each of the factors plays is not fully understood but the occurrence of organic and inorganic complexes is well documented (e.g., Williams, 1969). Wagner (1969) reviews the literature pertaining to the composition of the dissolved organic compounds in sea water but points out that “the most efficient procedures for separation and isolation tend to be also the best procedures for formation of artefacts”. The transport, solubility, adsorption, as well as biological uptake of trace metals in the sea are strongly influenced by the formation of metallo-organic complexes (Richards, 1965; Desai & Ganguly, 1970; Salo & Saxen, 1974). Sugimura et al. (1978) found that most of the dissolved copper was associated with organics, regardless of depth while Kerr & Quinn (1980) not only found that organic matter from different water masses exhibited varying spectroscopic characteristics but also that it varied in its ability to interact with copper ions. This is not surprising because of the effect of circulation as well as the

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unique biological nature of individual water masses and the effects of vertical movement of participates (e.g., Duce et al., 1972). Organic complexes may also adsorb onto clay minerals and be transported to the sediment or may accumulate at the air-water interface, become aggregated or enter the food chain. Riley & Taylor (1969) used Amberlite XAD-1 resin to recover organic materials from sea water, including humics with known complexing ability. Sephadex (Gjessing, 1965) and ultrafiltration (Gjessing, 1970) have been used for the estimation of the molecular weight of humics in fresh water. Chau & Chan (1974) examined the effect of some of these organics when they used ASV (anodic stripping voltammetry) to determine labile and strongly bound zinc, cadmium, lead, and copper in lake water. Maeda & Tanaka (1977) examined the various forms of copper in Tokyo Bay, separating the particulate from the dissolved phases and recognizing dissolved reactive copper and four groups of “total soluble” copper, through the results of various sample treatments. Huizenga & Kester (1979) isolated marine dissolved organic material by activated charcoal and examined the protonation equilibria with the assumption that the characterization of acid-base equilibria is required for the assessment of possible trace metal-organic interactions in sea water. Schmidt et al. (1978) characterized soluble copper species in estuarine sediments by means of ultrafiltration and then applied gel permeation chromatography (GPC) to some of the molecular weight fractions. A GPC fraction that appeared to contain aromatic and acidic components was subjected to thin layer electrophoresis which allowed separation of three mobile compounds (two positively and one negatively charged). The various organic substances in water often directly influence the chemistry of inorganic dissolved substances, especially trace metals (Krauskopf, 1956). Strong association of transition metals with organic substances within living systems makes it plausible that similar associations exist in natural waters in the form of metal-organic complexes in solution (Malcolm, Jenne & McKinley, 1970). Chlorophyll, for example, is the most commonly identified natural complexing compound in natural waters (Cline & Holland, 1977). Even although complexing agents have been suggested as the mechanism controlling the biological availability of copper (e.g., Schmidt, 1978a,b) and have been isolated from marine environments (e.g., Sieburth, 1971; Mantoura & Riley, 1975) Duursma & Sevenhuysen (1966) suggest that they are at low levels. They feel that it is easier to explain the occurrence of metal concentrations higher than accounted for by solubilityproduct values on the basis of the formation of basic cation complexes. Furthermore, they suggest that although organic compounds like isoleucine and arabinose do occur and could complex metals, their low concentrations prevent their being important mechanisms. This earlier work was supported by Zirino & Healy (1970) and Zirino & Yamamoto (1972) (see also Pocklington, 1977). The importance of inorganic complexes cannot be over-emphasized although it seems likely that the biological availability of copper is affected by organics as well as inorganics, by adsorption as well as complexation. In a study of the effect of physiochemical form of trace metals on their accumulation by bivalve molluscs Harrison (1979b) states that in the model used to assess environmental impact of power plant sitings the use of a single maximum concentration factor for bivalve molluscs in this situation is appropriate for screening purposes. She also states that “…when more realistic estimates are required, selection of

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concentration factors applicable to the site, species, and situation is necessary”. In one of the older studies on the absorption of copper, White & Thomas (1912) noted that Fundulus heteroclitus takes up copper through the gills, from a copper sulphate solution. Delisle et al. (1975) found that, in both goldfish and catfish, accumulation was proportional to the initial abundance of metals (zinc, lead, copper, cadmium) in sediments although the rate and degree of accumulation was stated to be due to chemical form, sediment particle size, agitation of the sediment, feeding habits, and physiological characteristics of the organism as well as the concentration of metal. With the knowledge that copper is an essential micro-nutrient (see Bowen, 1966; Manahan & Smith, 1973), the problem is to relate the geochemical events to the biologically available species of the metal. The dissolved phase appears to be biologically more available than the particulate (e.g., Lunz, 1972) and the ionic form more available than other dissolved species (Steemann Nielsen & Wium-Anderson, 1970). Cupric ion activity (pCu, the negative log of the cupric ion concentration) has been used to estimate the toxicity of copper, in marine and fresh waters as well as artificial media, to bacteria, phytoplankton, and fish (Sunda & Guillard, 1976; Anderson & Morel, 1978; Sunda & Lewis, 1978; Engel & Sunda, 1979). Plots of cupric ion activity against growth rates of phytoplankton normally show some level below and above which growth rates are reduced. The implication is that the organism requires a certain level (or activity) but that, when this is exceeded, the excess produces a toxic condition. It is also important to note that the pCu levels found to be optimal for many organisms are frequently exceeded in the natural environment (see Sunda & Hanson, 1979). In a study of Gonyaulax tamarensis, the organism producing red tides, Anderson & Morel (1979) indicate that the trace metal chemistry of the environment is important and that the complexation of toxic metal ions (e.g., copper) or solubilization of limiting metal ions (e.g., iron) limits the success of the population. Anderson, Morel & Guillard (1978) also found that low zinc ion activity can be a growth-limiting factor for Thalassiosira weissflogii and that it is the sum of the interaction of metals and complexing agents that is important. Cupric ion activity has also been shown to be important in the metabolism of a marine bacterium (Sunda & Gillespie, 1979) and with the copper sensitive bacterium, Klebsiella aerogenes, growth and survival were inhibited in the range of 10−8 to 10−6 M Cu2+ (Zevenhuizen et al., 1979). Dodge & Theis (1979) found that the freshwater larva of Chironomus tentans takes up copper from a solution in which the free cupric ion and a copper hydroxy-complex were dominant but that there was no uptake when the metal was in an organic complex. It should be pointed out that even although the cupric ion has been shown to be important, other species of dissolved copper are also of importance either directly or indirectly. Luoma (1978a) points out that increases in humate-bound copper increases the biological availability to two clams (Scrobicularia plana and Macoma balthica) in estuarine sediments. This may be due to the removal of sediment adsorbed copper by humates. In freshwater algae, copper compounds with co-ordination bonding are more toxic than compounds with co-ordination-covalent bonding (Khobot’yev et al., 1975). The complicated nature of the mechanisms involved in controlling the toxicity of copper can be exemplified by the freshwater study of Swader & Chan (1975) in which copper solubility and cupric ion toxicity in bicarbonate solutions was enhanced by the addition

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of citric acid. Laube, Ramamoorthy & Kushner (1979) used cultures of two freshwater algae, Anabaena 7120 and Ankistrodesmus braunii, to suggest that cadmium and copper present in sediments can be accumulated by algae and to suggest that algae may play an important rôle in mobilizing sediment-bound heavy metal ions. Neff et al. (1978), on the other hand, found that statistically significant accumulation of metals by several benthic invertebrates, from sediments in contaminated areas, occurred only slightly more than 25% of the time. One thing that emerges from the literature on cupric ion activity and the toxic effect of various species of copper is the suggestion that the equilibrium between the various species is important, that the concentration of the various species is controlled by a number of factors (see Orlob, Hrovat & Harrison, 1978). Long & Angino (1977), using a theoretical model for freshwater–sea-water mixtures found that the absolute and relative concentrations of the competing inorganic ligands were the major controls of the amount and type of inorganic complexing taking place in various natural water solutions. They tend to discount the importance of free amino acids and hydrocarboxylic acids as complexing agents. Metal speciation has been examined on theoretical grounds (e.g., Singer, 1977), for multiligand media such as sea water and concentrated brines (Van Luik & Jurinak, 1979). MacCarthy & Smith (1979) provide a quantitative model for complexation in multiligand mixtures and comment that existing models for computing equilibrium concentrations of all species in multimetal-multiligand mixtures work only when the stoichiometric concentrations of all metals and ligands, in addition to the stability constants for all complex species, are known, a condition not found in most natural systems! Ball, Jenne & Nordstrom (1979) describe a computerized chemical model for trace and major element speciation and mineral equilibria for natural waters. This model, WATEQ2A, includes copper and composite ligand groups (fulvates and humates) but defaults to molecular weights of 650 and 2000 for the two ligand groups. Jackson & Morgan (1978) used a theoretical model to analyse previously published data and showed a correlation between growth rate of phytoplankton and cupric ion concentrations, a relationship which Sunda, Barber & Huntsman (1981) have indicated could be partially due to the competition between copper and manganese at manganese nutritional sites. Earlier, Barber (1973) found that both Mn and Fe partially reversed the detrimental effect of U.V. photo-oxidation of natural sea water; this photo-oxidation decomposes organic complexing agents, releasing bound copper and increasing the cupric ion concentration. Leckie & Davis (1979) provide a good review of co-ordination chemistry of copper and speciation of copper in aqueous solution. They then apply this to a discussion of copper in natural aquatic systems. Lion & Leckie (1979) used equilibrium model computations to describe copper stability with inorganic and organic ligands typical of marine microlayer environments in order to attempt an understanding of these complex interfacial matrices. In an article on copper in natural waters, however, Boyle (1979) points out that “it is extremely unlikely that solubility equilibrium plays any significant role in determining the concentrations of copper in most natural waters. In the ocean, the biological and adsorption removal processes, and the atmospheric, river, and bottom input fluxes, combine with the oceanic circulation to exert a dynamic control on the copper concentration variations throughout.” In a publication on chemical speciation in river water, Sunda & Hanson (1979) point

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out that, based on modelling studies, the speciation of copper in the world’s rivers will be dominated by complexes with natural organic ligands, and that this should have a marked influence on both the biological and geochemical reactivity of the copper. Again in freshwater, Wilson (1978) uses an equilibrium model to describe the influence of humic materials on speciation of copper, zinc, and manganese and notes that the amount of metal bound by humic substances may control micro-nutrient availability. Chemical speciation, either calculated or measured, can be used to assess potential toxicity or copper complexing capacity of the water. Van den Berg & Kramer (1979) used MnO2 as a weak ion exchanger, to assess the complexing capacities and conditional stability constants for compounds in natural fresh waters. Based upon the calculation of the complexation of copper by 10−6 M naturally occurring ligands at different pH levels, they estimated that the cupric ion concentration is lowered between pH 5 and 9. An area of increasing interest is the specific nature of the factors and agents controlling the biological effect of copper. In a discussion of complexing metals by soil organic matter Mortensen (1963) points out the importance of identifying and characterizing metallo-organic matter complexes before unequivocal evidence for their formation in soils can be obtained. This philosophy is still applicable to today’s studies of estuarine and marine environments. Montgomery & Santiago (1978) state that the true chemical form of the metals should be analysed in order to understand the biogeochemical pathways of trace metals. They note that, in an estuary, the “soluble” organic ligand fraction can exceed 50% of the total “soluble” metal. Application of the techniques and concepts for understanding organometallic interactions in the soils (e.g., Schnitzer & Hansen, 1970) to fresh water (e.g., Gjessing, 1976) and marine environments (e.g., Bojanowski, Pempkowiak & Kupryszewski, 1977) will assist in developing an understanding of organo-metallic compounds necessary for application to determining the biological availability of copper and other metals in the marine environment. Means, Crerar & Amster (1977) found that gel filtration chromatography (GFC) could be used in fresh water to discriminate between different molecular weight organic fractions in terms of their ability to mobilize trace metals. Blutstein & Smith (1978), using ultraviolet light photo-oxidation techniques and anodic stripping voltammetry, demonstrated that most of the copper in the top metre of water in-the Yarra River Estuary (Australia) was associated with organics. Dehlinger et al. (1974) noted that a large fraction of the copper carried into Long Island Sound by the Connecticut River appeared to be organically bound. Sunda & Lewis (1978), used a mixture of 95% river water and 5% “Gulf Stream seawater” to demonstrate that something in the water could reduce the effect of copper on the division rate of the unicellular alga, Monochrysis lutheri. Whitfield & Lewis (1976) related the biological availability of copper (to a calanoid copepod) to organic material in sea water and to hydrographic and biological conditions. Barber, Dugdale, MacIsaac & Smith (1971), Barber & Ryther (1969), and Barber & Huntsman (1978) associated the effect of copper on the growth of marine phytoplankton with organics in the water and related their presence to hydrographic conditions. Gillespie & Vaccaro (1978) developed a bacterial bioassay to measure copper-chelation capacities of sea water and found that the capacities ranged from 3–40 µg copper·1−1. The change in the species of copper in water and sediments is due to a number of processes. Pellenbarg & Church (1979) note that litter from the salt marsh grass Spartina

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is involved with a microlayer on the surface of the water, that it both absorbs surface microlayer components and releases organic materials which form part of the microlayer. These organic materials could “provide natural soluble chelators for trace metals and account in part for net export as dissolved components”. They also note that the surface microlayer in the Delaware salt marsh carries an average of 10% of the copper relative to the total metal flux. Elzerman & Armstrong (1979) and Elzerman et al. (1979) found copper enrichment of the surface microlayer in fresh water. Vertical movement, either by bubbles (Wallace & Duce, 1978b) or by biogenic processes (Hamilton-Taylor, 1979) involve a change in both the physical state, from dissolved to particulate, as well as the chemical state. Naidu, Feder & Norrell (1978) examined the effects of Prudhoe Bay crude oil on a tidal flat ecosystem in Port Valdez, Alaska and found that the oil changed the sediment pH and Eh relationships enough to mobilize heavy metals from the sediments. They demonstrated a significant difference between the concentrations of copper, zinc, nickel, and vanadium in the oiled and control sediment samples which they attributed to the effect of the oil. Numerous authors (e.g., Jones, 1964; Lakshmanan & Krishnan Nambisan, 1977; Morris & Russell, 1973) have noted the effect of complexation on the biological effect of copper. Gächter et al. (1978), using 0.45 µm filtered lake water, found that material retained on an Amicon UM-2 membrane (retention of material with nominal mol. wt>1000) could regulate copper availability to phytoplankton. Gnassia-Barelli et al. (1978) cultured Cricosphaera elongata in a nutrient-enriched medium without complexing agents and found that roughly half the copper added (50–100 µg·1−1) was complexed in the culture filtrate. (Stolzberg & Rosin, 1977, found similar values with Skeletonema costatum.) Gnassia-Barelli et al. (1978) then filtered the medium with an Amicon ultrafilter and, from the distribution of the copper, concluded that the complexing agents in the medium were mainly those of a mol. wt ranging between 500 and 10 000. The relationship between copper-complexing capacity and copper toxicity has been examined for the zoeal larval stage of Pandalus danae (J.S.Young et al., 1979a) and for the juvenile Atlantic salmon Salmo salar in fresh water, in the presence of humic acid (Zitko, Carson & Carson, 1973). In both cases, toxicity could be related to the level of the AS V-labile metal. Hutchinson & Collins (1978) discuss metal species in soils, fresh water, and sea water and the effect of hydrogen ion activity and calcium on the toxicity of metals in the environment. The chemical process of complexation is a quantum mechanical-interaction. An electron donor forms a ligand for an electron acceptor, which is the metal. The general form of this reaction, where n is the co-ordination number of the metal ion, Me, the metal, m, the oxidation state of the metal, L, the ligand, and q, the ionic form of the ligand, is:

The nature of the mechanism is discussed at length in several recent publications (Levine, 1979; Singer, 1974, 1977). Organic functional groups that are likely to be involved in binding copper to organic compounds are carboxylate, enolate alkoxide, phenoxide, aliphatic, amino, aromatic amino, mercaptide, phosphate, and phosphonate (Martell,

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1971). Thus, nearly every biological molecule is capable of participating in the complexation of metals. The number of ligand groups on the organic molecule as well as the configuration determines how the organic occupies the co-ordination sites of the metal atom. Properties of the organic ligand which may affect the stability of metal complexes are structure, ability to form chelates, polarizability, and types of available binding atoms (Ahrens, 1966; Lehman, 1963). The term chelation is applied to the condition in which one molecule of an organic is able to occupy more than one of the co-ordination sites of the metal. This is essentially the formation of a ring complex about the metal atom. Chelating agents have a set of common properties and each metal may exhibit a preferential bonding to certain ligand groups. All chelate rings include the metal in a five-membered ring which provides stability. As an indication of the nature of the process in the natural environment, Hallberg (1978) examined metal-organic interactions at the sediment redoxcline and comments that “in sediments the microbiological decomposition processes of organic matter are nonsynchronized and result in an increase of intermediate compounds, which may act as metal chelates. Therefore, there will be a competition for the heavy metals between the chelating agents and the hydrogen sulfide.” Organics capable of forming organometallics often occur as polymers or are joined by means of oxygen and nitrogen functional groups in a six-membered ring called a clathrate compound (Degens, 1970). These large organic molecules often display colloidal properties in solution (Ogden, 1920) such that changes in ionic strength, presence of heavy metal cations or changes in pH can cause precipitation of a low specific gravity floc (Breger, 1970). A number of techniques have been used in an attempt to differentiate routinely between the various forms of copper. Anodic stripping voltammetry (ASV) has been used to differentiate between strongly and weakly complexed copper. Gurtisen, Crecelius, Abel & Philbrick (1978) found that the copper complexation capacity, as measured by ASV, increased from the surface to near the bottom in coastal waters, ranging from 3–45 µg copper·1−1. ASV titration of marine humic-copper complexes yielded stability constants of 105–106 (Lieberman & Healy, 1978). Complexation capacities of various molecular weight fractions (Amicon filtration separations) of dissolved organic matter in estuarine waters have also been examined by ASV (Smith, 1976). Smith found that the copper complexation capacity of the lower molecular weight fractions appeared to increase with increasing salinity suggesting a marine source of low molecular weight materials. The cupric ion electrode has also been used to estimate cupric ion activity with respect to complexing agents. Williams & Baldwin (1976) noted that “complexation of copper with organic matter appears to be the major

mechanism preventing high cupric ion activities which, otherwise, could be toxic to marine phytoplankters”. McKnight (1978a, b) and McKnight & Morel (1979) examined the ability of a number of algal species, including four estuarine eucaryotes, to produce copper-complexing agents. All four of the estuarine species produced measurable amounts when grown in a freshwater medium, as detected by potentiometric titration (cupric-ion electrode), agents that acted like weak organic acids. Several of the

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freshwater blue-green algae tested were found to produce much stronger copper-complexing agents that gave positive Csaky test reactions indicating the presence of bound hydroxamates, indicative of siderophorelike iron-complexing agents (e.g., Neilands, 1974). Three chlorophytes and two cyanophytes released copper-complexing agents during the stationary phase of growth (senescence) but not during the exponential phase (rapid growth). Williams & Rosen (1978), using the cupric ion electrode, tested a series of amino acids, Krebs cycle acids, uronic acids, and fulvic acids isolated from surface sea water. Copper chelation was dependent upon the inherent stability constants of the complexes. Histidine was the only ligand of those tested which complexed the cupric ion to any extent in sea water. The determination of metal species is made by the measurement of metal levels with techniques that estimate only a certain fraction of the metal, or after treatment with a technique that will change the state of certain fractions of the metal. Each of these techniques has certain inherent problems which makes the determination of metal levels and of each species difficult (e.g., Duinker, 1978). Zirino & Kounaves (1980) discuss the problems of using ASV for the determination of copper in sea water and point out that acidification and/or the addition of ethylenediamine will improve the method. These techniques however, change the nature of the metal species. Heuss & Lieser (1979a) discuss the problems associated with neutron activation analysis of trace metal levels in sea water and discuss (Heuss & Lieser, 1979b) a technique for adsorption of copper on activated charcoal pointing out that decay may occur which would change the nature of the copper species. Westall, Morel & Hume (1979) examine chloride interference in cupric ion selective electrode measurements and suggest that the chloride ion effect accounts for the general unsuitability of the electrode for measurement of cupric ion activity in sea water. This paper is important because the cupric ion has been shown to be biologically important and direct measurement would be ideal, as is possible in fresh water (e.g., McCrady & Chapman, 1979), as it would allow rapid estimation of one of the copper species that is biologically available. With all of the problems of analysis, even general information on the speciation of copper in estuarine and marine environments is of real importance in determining the biological impact of the metal. In contrast to Orren (1979) it is important to realize that to better understand the natural complexes that exist it is necessary to attempt to characterize them, first in a general manner (Batley & Florence, 1976b) and then with greater precision. One group of organic agents that has long been considered to be important in controlling the biological availability of metals, including copper, is the large array of chemicals called “humic” and “fulvic” compounds (see Rashid, 1971; Manning & Ramamoorthy, 1973; Rashid & Leonard, 1973; MacFarlane, 1978). Humic acids are defined as that group of compounds which can be extracted from soil or sediment by alkaline solution and then precipitated upon acidification, while fulvic acids are those compounds that can be extracted by alkaline solution, but remain in solution upon acidification (see Kononova, 1966). A key feature of both compounds is their low solubility in water and the large number of active sites. Morris & Calvert (1977), in a

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geochemical study of organic-rich sediments from a core from the Namibian shelf, found fulvic acid levels to be high and humic acid levels lower. They suggest that conversion of the planktonic organic material to the higher molecular weight humic compounds appears to be slow. Bojanowski et al. (1977) isolated and obtained preliminary characterization of humic substances in Baltic sediments and suggest that heavy metals may provide a stability to the complex. Peake (1978) found humic and fulvic acids in sediments from Nancy Sound in New Zealand while Simoneit et al. (1979) found humate and kerogen concentrations estimated from 9–19 mg·g−1 dry sediment in a 3.5 m gravity core from an area of high geothermal gradient in the Guaymas Basin (Gulf of California). Stuermer & Harvey (1977) isolated humic substances from sea water from both nearshore and oceanic locations off the east coast of the U.S.A. and found that the ratio of humic to fulvic acids varied suggesting a localized effect either on the source or the conversion of organic material to humics in the water column. MacFarlane (1978) isolated these compounds from the sediments of a northern Florida estuary and found the humic acids to be of a molecular weight range of <500–>300 000 with the greatest concentration at the higher end. The fulvic acids were most commonly found at <500 and in the 100 000– 300 000 molecular weight range. The effect of humic compounds in fresh water is to solubilize metals (e.g., Buffle, Greter & Haerdi, 1977) and to mediate metal-ion transport (Pagenkopf, 1978) in rivers. Prakash & Rashid (1968) were among the first to point out the importance of humic compounds in controlling the biological availability of trace metals in marine waters. Hopner & Orliczek (1978) examine humic material as a component of sediments in estuaries, based on the observation that dissolved and colloidal humic material entering an estuary is precipitated as a result of mixing with sea water. They comment that “… humic matter in a sediment may have important functions in the ecology of an estuary and its sediment, binding toxic heavy metals on its hydrophilic shell, or hydrophobic pollutants in its hydrophobic core. Apart from this, humic matter seems to be one of the final forms of carbon deposition in principle re-utilizable—but if so, only extremely slowly.” Rashid & King (1969) used sephadex gels to estimate the molecular weight distribution of humic and fulvic acid fractions from marine clays on the Scotian Shelf. They found values ranging from 700–>2 000 000 mol. wt. Rashid & King (1971) also chemically characterized sephadex fractionated humic acids associated with marine sediments and found, among other things, that C/H ratios increase with an increase in molecular weight. Steurmer & Harvey (1974) used sephadex to estimate the molecular weight of humic substances isolated by XAD-2 resin from sea water. Sephadex to estimate the molecular weight of humic substances had been used earlier by a number of workers (e.g., Gjessing, 1965; Gjessing & Lee, 1967) and has shown that organic carbon elution patterns varied depending on the origin of the material. Stuermer & Payne (1976) compared humic substances from sea water and terrestrial sources with carbon-13 and proton nuclear magnetic resonance and found structural differences which seemed to result mainly from the low abundance of aromatic precursors in sea water when compared with terrestrial material. They comment that “the incorporation of marine lipids and pigments into the products of the Browning reaction could account for the structural features indicated by the present work”. Rashid (1971) found that the lower molecular weight humic fractions complexed 2–6 times more metal

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than the higher molecular weight fractions. Wilson & Kinney (1977) noted that humic materials have many of the chemical and physical characteristics of enzymes which they felt could allow the use of modified equations from the biochemical literature to provide a theoretical description of proton-metal ion interactions. One of the things noted with respect to the occurrences of red tides has been the association of outbreaks with increased land drainage. Whether this is due to the increase in complexed iron (Martin, Doig & Pierce, 1971), a combination of complexing agents, nutrients, and pollutants (Prakash, 1975) or the effect of run-off-produced complexing agents on cupric ion activity (Anderson & Morel, 1978) is open to debate. In all probability it is a combination of these factors. Mantoura et al. (1978) isolated humic materials from sea water, river water, and lake water, and worked out the stability constraints with a number of metals. At a pH of 8.0 the order of increasing strength of binding of the metals generally followed the Irving-Williams series. In fresh water >90% of the copper was complexed by humic materials while in sea water only ≈10% was complexed. Mantoura et al. (1978) found that up to 80% of the metal may be complexed. In reporting a gel filtration study of trace metal-fulvic acid solutions from sewage sludges, Baham, Ball & Sposito (1978) suggested that copper and iron formed complexes with sludge-derived fulvic acid and that the acid behaved as a relatively harder Lewis base than Cl−. A study of anaerobically digested sludge (Gould & Genetell, 1978) gave similar results. In a study of the rôle of humic acids during transport of trace elements in the marine biocycle, Huljev & Strohal (1972) found that degradation of humic acids occurs by the action of micro-organisms and that during this process certain trace metals become available to marine animals. Juste, Delas & Langon (1975) noted, however, that the addition of polyvalent metals to humic acids considerably diminished the biodegradability of the humates. They stated that this addition blocked the reactive sites of the organic material and had a toxic effect on the microbial flora. The suggestion was that the humic acid “phase” of the marine biocycle would be extended if polyvalent metals were complexed. This has been found for other complexing agents; Walker (1976) noted that biodegradation of nitrilotriacetate (NTA) in fresh water was inhibited by copper and cadmium at metal to NTA ratios of 1:1. The inhibitory effect of both metals was reduced by the presence of iron and the inhibitory effect of copper by water hardness. Seeliger & Edwards (1979) found that up to 22% of the copper in the thalli of two benthic red marine algae was associated with organics released as dissolved material after metal uptake. After decomposition of the thalli, 80–90% of the copper was associated with dissolved or particulate organic material. Rashid & Prakash (1972) examined exudates of decomposed thalli of Fucus vesiculosus and Laminaria digitata and found 20–40% of the humic acid compounds represented by fractions having a molecular weight >200 000 whereas the fulvic acids did not exceed 10 000. They found the dominant functional groups to be carboxyl and carbonyl but with-phenolic and alcoholic hydroxyl groups also present in small proportions. 13C and proton NMR spectra for two representative fulvic acid samples showed significant amounts of aliphatic and aromatic constituents including products from degradation of polysaccharides and proteins (Sposito, Schaumberg, Perk & Holtzclaw, 1978). Analyses by Stuermer & Harvey (1978) concur.

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Sieburth (1971) demonstrated a loss of humic materials in estuaries and indicated that trace metals would be removed along with the humics. Pellenbarg (1978) found that Spartina intercepted trace metals carried as dissolved and particulate burdens in the water, especially as part of the surface microlayer. The result was copper-enrichment of Spartina litter of some 2.5 times. Spartina also released substances during decomposition that behaved like fulvic acids. The concentration of trace metals and the production of fulvic acid-like materials may form a mechanism for introducing trace metals to the biota of a salt marsh. It may also form a near-marine source of these materials. One of the sources for organometallic copper introduced into the marine environment is sewage. Bergman et al. (1979) for example, found that most of the trace elements in sewage sludges were located in organic compounds and that particular organic compounds contained the greatest concentrations of trace elements. The nature of the organic compounds, as well as their concentration has been found to vary with the nature of the sewage plant operating conditions as well as the type of material sent to the plant (Katz, Pitt, Scott & Rosen, 1972). Copper-binding agents have been found in a number of marine and non-marine organisms. Naiki & Yamagata (1976) isolated a copper-thionein-like protein (mol. wt ≈ 10 000) from a copper resistant strain of Saccharomyces cerevisiae. Soil bacteria have been found to produce polysaccharides capable of complexing with metals (Lasik & Gordiyenko, 1977). Bere & Helene (1979) discuss the binding of copper to polypeptides containing glutamic acid and tyrosine residues, as indicated by a strong absorption band around 245 nm. Briand, Trucco & Ramamoorthy (1978) used ion specific electrodes to determine heavy metal binding in lakes, and correlated it with specific algae. Binding capacity for copper, mercury, lead, and cadmium tended to be associated with certain species of algae that made up only a small proportion of the total algal volume. Howard & Nickless (1978) found a low molecular weight (<3000) water-soluble extract in each of three species of molluscs that appeared to be able to complex copper. Coombs (1974), working with Ostrea edulis, found that ≈40% of the copper and zinc was weaklycomplexed to compounds with small molecular weights, such as taurine, glysine, ATP, and possibly homarine (N-methyl-alpha-picolinic acid). These complexes can act as a freely available source of metal for metabolic requirements. Howard & Nickless (1977b) were unable to find metallothionein-type proteins in O. edulis and Crassostrea gigas from polluted habitats but did find that they contained three distinct low molecular weight zinc complexes, only one of which was present in individuals from less contaminated conditions. As with Coombs (1974) and Howard & Nickless (1975) copper appeared to be associated with amino acids, principally taurine, and the betaine homarine. Copper may also be concentrated in extracellular polymers from marine bacteria (Corpe, 1975). These polymers have been found to form insoluble precipitates with cationic detergents and soluble salts of iron, copper, and lead. Eichhorn (1975) provides a detailed discussion of organic ligands contained in living organisms for metal binding. Among these he lists nucleic acids and their constituents (nucleotides, nucleosides, and phosphates), amino acids, and peptides which bind metals through carboxyl and amino groups as well as through the functional groups that characterize specific amino acids. For example, SH− in cysteine, CH3S− in methionine, OH− in serine, imadazole in histidine, and NH2 − in lysine. Eichhorn also notes low molecular weight oligopeptides

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which have been designed in nature for the specific purpose of complexing metal ions. Oligopeptides are the ‘ionophores’ which bind alkali metals and carry them across lipid barriers. They generally form rings in which the metal ions are combined. Not all are peptides but those non-peptides are somewhat similar in structure to the peptide ionophores. Eichhorn also discusses proteins which are active in metal storage and transport via their functional groups, e.g., ceruloplasmin (Lontie & Witters, 1973) which is specific for copper, and haemocyanin (Scheinberg & Morell, 1973). Eichhorn also mentions metallo- and co-enzymes, porphyrins, haemoproteins, carbohydrates (which contain hydroxyl groups) and phospholipids as potential metal-binding agents. Gelbstoff, a lignin-type material, also appears to play a dominant rôle in organo-copper interactions (Schmidt, 1978a,b). Khailov (1964) examined a number of marine algae from the Black Sea and found indications of complexation of copper by the external metabolites of at least one, Dunaliella salina. Swallow et al. (1978) found that the freshwater diatom Gloeocystis gigas produced an exudate capable of copper complexation and McKnight & Morel (1980) have found that cultures of the blue-green algae Anabaena flos-awuae and A. cylindrica release an organic, which they term a siderophore, capable of complexing copper. They suggest that the siderophore-like agent(s) may reduce the toxicity of copper because the organo-metallic complex is less toxic than ionic copper. Andersen, Le Blanc & Sum (1980) isolated and characterized an extracellular metabolite from Prorocentrum minimum which is a β-diketone with characteristics suggesting the possibility of copper complexation. Burnison (1978) isolated high molecular weight (>300 000) polysaccharides from lake water by ultrafiltration and comments that “it has been speculated that they act as ion exchangers to transport nutrients to algal cells, or that they bind heavy metal ions and thereby protect the algae from toxic concentrations”. A variety of polysaccharides have been isolated from extracellular products of algal metabolism (e.g., Merz, Zehnpfennig & Klima, 1962) making this a potentially important source of complexing agents. Duursma (1970) found that the amino acid leucine could complex both cobalt and zinc, the only two metals tested. Armstrong & Van Baalen (1979) discuss the production of a siderophore by the marine blue-green alga Agmenellum quadruplicatum. Hydroxylamine production has also been noted for an actively nitrifying Arthrobacter sp. isolated from a lake (Berger, Rhô & Gunner, 1979), production which suppressed the growth of Chlorella vulgaris. Since hydroxylamine production is a result of trace metal deficiency, this suggests successful competition for trace metals by Arthrobacter. Davies (1970) discusses growth of Dunalliella tertiolecta under iron-limiting conditions and suggests that chlorophyll production may require the presence of organo-iron complexes. The organic complexing agents, whether amino acids or siderophores, may also form important agents for copper complexation. Howard & Nickless (1975, 1977b) found the major portion of copper in the molluscs Patella vulgata and P.intermedia to be associated with a protein with “a molecular weight of 10 800 daltons” which was similar to mammalian metallothionein. They attributed its high binding capacity to the high cysteine content of the molecule. Howard & Nickless (1978) also found accumulations of cadmium, zinc, and copper by several species of molluscs from polluted situations. Accumulated metals in the animals were associated with low molecular weight organics suggesting that complexation had

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occurred. Noel-Lambot (1976) found that zinc and copper were principally associated with high molecular weight proteins in Mytilus edulis as was cadmium under normal conditions. With cadmium stress, however, a low molecular weight protein, similar to metallothionein, was produced and was associated with the cadmium. Metallothioneinlike compounds have also been isolated from marine vertebrates by Olafson & Thompson (1974) who found them to be of molecular weights 9000, 10 000, and 11 000 in Atlantic grey seals, Pacific fur seals, and copper rock fish, respectively. Metallothionein, or analogues, have also been found in Scylla serrata (Olafson et al., 1979), Patella vulgata, Littorina littorea, and Purpura lapillus (Noel-Lambot et al., 1978a) as well as Mytilus edulis (Talbot & Magee, 1978) and Anguilla anguilla (Noel-Lambot et al., 1978b). The activity of these compounds is also discussed by Parsons & Brown (1978) and Brown, Bawden & Chatel (1977) for other marine species. Rosen & Williams (1978) state that “the speciation of cupric ion and in situ copper equilibrium is controlled by the inherent stability constants of the available ligands, the pH of the medium, and the rate of exchange between various copper species”. The nature of the relationship between the copper and the organics appears to change on a seasonal basis (Morris, 1974) and even in a diel pattern (Johnson, 1978). Johnson, using ASV, showed that the percentage of ‘free’ copper in unfiltered sea water was significantly lower in the daytime than at night. He points to the potential hazard of speciation analysis, suggesting that copper speciation may change over a short time, possibly due to differences in the uptake processes of phytoplankton during night and day. Add to this the findings of Zirino, Lieberman & Clavell (1978b) and Young, Adams & Darby (1977) that inshore copper concentration can vary on a tidal basis, and it becomes apparent that metal species and metal concentration, the two major factors that control the biological availability, can vary over short periods. One then wonders at the usefulness of ‘classical’ toxicity studies in relation to real life situations. Introducing a complexing agent into the metal-organism relationship may produce a complex system because of the interactions of the metal with natural adsorbing agents. Elliot & Huang (1979a) examined the adsorption characteristics of Cu(II) in the presence of chelating agents and found that the agents improved the extent of Cu(II) adsorption, at least in the Al2O3 system used. Davis & Leckie (1978a,b) found that complexing ligands can be important in determining trace metal adsorption on hydrous oxides in natural systems. Some ligands (e.g., picolinic acid) when adsorbed, cannot complex copper ions whereas others (e.g., glutamic acid) can. Metals and organics in interstitial water are derived from the sediments themselves or from organism decomposition (Duinker, Van Eck & Nolting, 1974; Sommer et al., 1976). Burnett (1975) states that, in sedimentary deposits in the Equatorial Pacific, “metals such as Cu and Ni probably enter the sediments as immobile metallo-organic complexes”. The rate of accumulation is controlled by the rate of accumulation of organic matter as well as by environmental conditions (Sommer et al., 1976). The amount of metal complexed by organics such as humic compounds in interstitial water is believed to be high, exceeding 80% (Mantoura et al., 1978). Lieberman & Healy (1978) note that humic substances from marine sediments have conditional stability constants of 105–106 for copper. Schmidt & Gibson (1978) examined the influence of organic matter on trace metal flux in coastal sediments and suggest that copper in the form solubilized by dilute acetic acid is the

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probable source of copper from amended sediments. This would be inorganically precipitated or weakly absorbed metal. Cline & Holland (1977) examined the reaction of heavy metals with pore-water organics and found that the cupric ion reacts strongly with the organics in a non-linear step function, forming a stable complex. The effect of anoxic conditions in sediments is to reduce the mobility of both iron and copper (Bonatti, Fisher, Joensuu & Rydell, 1971) by transforming them into sulphides. Bender & Klinkhammer (1978) found that in sulphate-reducing sediments, the formation of insoluble sulphides of copper, cadmium, and nickel produced concentrations in pore water equal to or less than those in near bottom water. They suggest that the release of metals from anoxic sediments is not an important factor in the balance of the metals in water overlying the sediment. Meyers & Quinn (1974) note that, since humic substances constitute the bulk of sediment organic material, they are probably responsible for the lower adsorption of copper than the authors found in natural sediments. Emerson & Harrison (1978) found that a decrease in ‘soluble’ (<0.45 µm) matter significantly increased the adsorption of 64Cu to particulates. PARTICULATE COPPER AND ADSORPTION Copper is found in association with particles of lithogenic and biogenic origin or forming a component of sea salt particles in the atmosphere. Demina & Fomina (1978) found that most of the iron, manganese, zinc, and copper in surface suspended material of the Pacific Ocean occurs in association with iron and manganese oxides although a considerable amount of the copper (average of 31.2%) is combined with organic matter. Run-off provides the major source for particulate copper in nearshore waters but once in the estuarine environment, copper and other metals are found in the water column, the bottom sediments, in the interstitial water, and on very fine fractions (possibly colloidal iron and manganese oxides) in the water above the sediments (Skei et al., 1973; Bricker, Ferguson & Huggett, 1974; Duinker et al., 1974; Müller & Fôrstner, 1975; Duinker & Nolting, 1976; Lu & Chen, 1977). Carritt & Goodgol (1954) discuss general implications of adsorption reactions and Huang (1977) discusses the general nature of adsorption and the interaction of organics and true metals with adsorbing surfaces. Batley & Florence (1976a) consider that the greatest fraction of labile copper is present as species adsorbed on organic colloids. Adsorption of dissolved trace metals by hydrous oxides may be an important estuarine process resulting in accumulation of metals in coastal sediments (Grieve & Fletcher, 1976) although both clays and hydrous oxides are important (Burrell, 1973; Payne & Pickering, 1975; Stern, 1975; Stoffers, Summerhayes, Förstner & Patchineelam, 1977; Batley & Gardner, 1978; Hunter, 1980). The relationship between particle size and metal concentration varies and is dependent upon the processes producing the association between the metal and the particle (Gupta & Chen, 1975; Bradford, 1976; Farrah & Pickering, 1977; Foster, Hunt & Morris, 1978; Montgomery & Santiago, 1978). The biological effect of particle formation is thought to change the biological availability of metals by reducing the concentration of the dissolved species (Babich & Stotzky, 1977; Pesch & Morgan, 1978). Uptake of copper, by particulates, produces a change in the physical as well as the chemical state of the metal. Adsorption is a mechanism that can act to control

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concentrations of soluble metal from the water column as well as from sediment extracts (e.g., Rendell, Batley & Cameron, 1980), either in the ionic form or associated with complexing agents (Davis & Leckie, 1978a,b). Babinets et al. (1979) examined the copper adsorption behaviour under the redox conditions of the Black Sea and estimated chemical forms of copper by thermodynamic calculations. Melluso, Di Filippo, Izzo & Paoletti (1978) suggest that the anomalous distribution of copper, lead, and chromium in samples of the tunicate Ciona intestinalis found in the Neopolitan harbour area is due to the sedimentation of particles to which the metals are adsorbed. Zsolnay (1979) examined seasonal variation and the possibility of river input of colloidal carbon into coastal regions. The study points out differences in colloidal material from river and from coastal water and suggests that flocculation occurs at the freshwater–salt-water interface which means that “any pollutant or naturally occurring compound that has been adsorbed onto the organic colloid material in the river will also be removed to the sediment and not be introduced into the ocean”. Zsolnay also found that marine colloidal material does not contain appreciable amounts of aromatic compounds which compares favourably with the statement by Stuermer & Payne (1976) for marine humic material. Bader, Hood & Smith (1960) examined the recovery of dissolved organic matter and organic adsorption by particulates formed from phytoplankton cultures to examine the origin, nature, and concentration of organic material dissolved in sea water and how this is incorporated into the bottom deposits. They comment that the mechanism may be important in the accumulation of metallic ions in sediments and found that the nature of the particles and the types of organic compounds present will be, in part, a controlling factor. Elliot & Huang (1979b) view the effects of complex formation on the adsorption characteristics of heavy metals and conclude that the important factor in determining the influence of complexation on adsorption is the ability of the complexes to bind to the surface, not the ability of the ligand or the metal by itself. Adsorption may occur not only with the ionic form of copper but also with organometallic compounds (Batley & Gardner, 1978; Davis & Leckie, 1978a). Davis & Leckie (1978a) suggest that particulate matter coated with adsorbed humic compounds may contribute to trace metal partitioning by aquatic sediments. Wind-generated foam on Lake Mendota was found to be enriched with particulate matter, including heavy metals such as copper (Eisenreich, Elzerman & Armstrong, 1978). The foam contained proteinaceous and carbonaceous material and had chemical properties similar to surface microlayers, which suggests that the particles were formed in situ. Foster & Morris (1971) indicate that the organically-bound copper in the marine environment is not controlled by land sources but is of marine origin. Adsorption of copper by biotic material is proposed as the most likely process for copper accumulation in particulate matter during plankton blooms (Abdullah & Royle, 1974). In general, organics appear to be important components of copper-containing particles. Spencer et al. (1978) found organically bound copper in faecal pellets taken from a sediment trap at 5367 m depth in the Sargasso Sea; Subramanian et al. (1974) found that chitin (from the cuticle of arthropods) forms a naturally occurring N-acetylated aminopolysaccharide that is capable of forming complexes with a number of metals, including copper. They indicate that the complexation can be described with adsorption isotherms and ion exchange mechanisms. The use of chitin as a metal-adsorbing agent

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has also been made by Hung & Han (1977). Corpe (1975) found that copper may be concentrated in extracellular polymers from marine bacteria. These polymers form insoluble precipitates with cationic detergents and soluble salts of iron, copper, and lead. In a study of heavy metal complexation behaviour in anaerobically digested sludges, Gould & Genetelli (1978) found that copper was the most strongly bound metal of those studied. The order of affinity of the metals for the sludge solids was on a moles bound per unit weight of solids basis. Transport of particulate copper to the bottom sediments may be either in the form of organic or inorganic material (e.g., Fowler, 1977). Spencer et al. (1978) suggested a fairly rapid transport of biogenic material from near the surface to the bottom in the Sargasso Sea. This means that there would be a net loss of copper from the water column and an increase in the bottom sediments. This was found to be affected by hydrographic and topographic conditions in a coastal situation (Lewis, 1979). The distribution and chemical partitioning of copper in benthic sediments is controlled by grain size as well as the chemistry of the sediments (Nissenbaum, 1974; Loring, 1976; Willey, 1976; Davis & Leckie, 1978a, b; Khalid, Patrick & Gambrell, 1978). Duinker et al. (1974) noted that copper, zinc, manganese, and iron were mobilized in bottom sediments, increasing the levels in the interstitial water in the Dutch Wadden Sea. They indicated that particular hydrographic and meteorological conditions may exist which produce exchange between interstitial water and overlying water and state that in the Wadden Sea the significance of metals, e.g. the uptake by bottom organisms, may be underestimated if only concentrations in the sediment are considered; that in the suspended matter should also be considered. In discussing the effect of dyking they stress that removal of the effect of tides from an area where exchange between interstitial water and overlying water prevails may cause high metal concentration in the interstitial water with great ecological significance. The redox potential and pH are two of the most important chemical properties controlling metal speciation within sediments (Khalid et al., 1978). In the study of a dredge-spoil area (Anderlini et al., 1975) and one on trace elements in Dead Sea sediments (Nissenbaum, 1974), it was apparent that copper frequently occurs as an organometallic complex. With the general tendency for copper in the sediments to be as a complex, the biological availability should be affected, an important consideration because the level of the metal is frequently high (e.g., Stephenson & Taylor, 1975; Luoma, 1978a). Luoma notes, however, that increases in humate-bound copper increase the biological availability of the metal suggesting either that the organisms are able to utilize the organometallic compound or that the equilibrium between the various copper species, under the effect of the humates, is to produce higher levels of biologically available copper. In a comparison of the bio-availability of several metals (not copper), Luoma & Jenne (1976) found that weak acids and weak reducing agents appeared to extract a different fraction of bound Cd, Co, and Zn than did the digestive processes of the clam Macoma balthica. Although relatively little work has been directed towards the particular rôle of ‘colloids’ in copper speciation and biological availability, numerous studies have indicated that they may be of major importance. Florence & Batley (1976), for example, suggested that some copper is adsorbed on, or occluded in, organic or inorganic colloidal

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particles. They also suggested that one of the effects of complexing agents may be the solubilization of iron which would have formed colloids capable of adsorbing other metals. This investigation, as well as an earlier one (Florence & Batley, 1975), examined the application of chelating resins to study metal speciation in sea water. Kremling & Petersen (1978) examined the relationship of particulate zinc, cadmium, and copper with iron, indicating that particulate iron was an important carrier for trace metals. Swallow & Morel (1978) noted the ability of ferric oxide to adsorb copper and attempted to relate this to a number of variables. FORMATION OF PARTICULATE METALS The change from the dissolved phase to the particulate can occur through any of a number of mechanisms (see Goldberg, 1954; Schmidt, 1978a,b). Several of the processes that affect metal speciation are associated with estuarine conditions (e.g., Demina, Gordeyev & Fomina, 1978; Burton, 1978a; UNESCO, 1979) and have been discussed under sources of copper (pp. 522–537). Duinker & Nolting (1978) found precipitation of copper dissolved in River Rhine water as a result of estuarine mixing and noted that trace metal concentrations in bottom sediments of the Rhine Estuary reflect the mixing and removal processes occurring in the water column. Morris et al. (1978) note the importance of the freshwater-seawater interphase and point out that the estuarine chemistry of Mn, Zn, and Cu may be different. Eaton (1979a) examined the relationship between soluble copper and river discharge in the San Francisco Bay and (Eaton, 1979b) notes that “the behavior of Cu and Zn is affected not only by the presence of nonriverine sources but also through the surface-active processes and organic complexing, coupled with the transportation of fine-grained sediments”. Förstner, Müller & Staffers (1978) discuss sources, chemical associations, and diagenetic effects of heavy metal contamination in estuarine and coastal sediments, noting a loss of soluble copper at the river-sea interface, and a tendency for re-mobilization from sediments due to a salinity increase, lowering of pH, and input of organic degradation products as well as synthetic complexing agents, change in redox conditions, and microbial activity. Differences in the concentration of trace metals through the water column may be due to uptake of metals by phytoplankton as well as adsorption by hydrous oxides (Abdullah & Royle, 1974; Bewers, Sundby & Yeats, 1976; Dehlinger et al., 1974; Krauskopf, 1956; Kremling & Petersen, 1978; Helz, Huggett & Hill, 1975; Topping & Windom, 1977) although there is some question about the nature of the exchange mechanism involved (Kharkar, Turekian & Bertine, 1968; Murray & Murray, 1973). Topping (1974) calculated the metal taken up by phytoplankton as 3.3P×CF/104B where P is the primary productivity (gC·m−2·time−1), CF is the concentration factor of the metal in sea water (µg·m1−1), and B is the solid (particle). Although the adsorption of copper to clay minerals is known throughout the marine environment (Goldberg, 1965; Carritt & Goodgal, 1954) the effect is probably greatest in inshore regions of high run-off. The reactions which may occur in the marine environment, with particulates are: (1) adsorption, (solid); and (2) ion exchange,

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(aqueous); where (aqueous=‘dissolved’, and A=exchangeable cation.) Because clay minerals enter the sea from the continents and only the finer particles remain in suspension, there is enough time for reactions of these types to occur. The concentration of copper is also minute when compared with some of the other metals also involved in this type of reaction. As a result, the actual amount of copper in deep-sea sediments (in which the clay minerals are a significant component) is relatively small although it can be predicted from its solubility (e.g., Krauskopf, 1956). Copper adsorption by hydrous alumino-silicate clay minerals is variable and dependent on the species of clay, pH, concentration of competing cations, as well as the nature and concentration of any ligands present (Farrah & Pickering, 1977). A number of clay minerals have been investigated; among the more common are montmorillonite, kaolinite, illite, and halloysite (Upadaya, Biddappa & Perur, 1974; Babick & Stotzky, 1977; Farrah & Pickering, 1977). Farrah & Pickering (1977) show that the adsorption of heavy metals increases with pH to a threshold for formation of sparingly soluble hydroxy-complexes. They also give the affinity order to divalent cations and show the nature of the adsorption processes to vary with clay type: illite

kaolinite

montmorillinite

The uptake of copper from solution by Fe- and Mn-oxide is suggested as a probable mechanism in coastal areas (Parks, 1967). This rôle is influenced by several factors including Eh, pH, copper concentration, concentrations of competing metals, and organic chelating agents (Jenne, 1968). The uptake mechanisms are co-precipitation (Foster & Hunt, 1975), surface complex formation, and ion exchange (Jenne, 1968; Stumm & Morgan, 1970). Copper adsorption by these oxides is widely documented in the literature (Batley & Florence, 1976a; Lopez & Lee, 1977; Davis & Leckie, 1978a,b; Rosen & Williams, 1978; Vuceta & Mortan, 1978). Krauskopf (1956) demonstrated an uptake of over 96% of existing copper by hydrated MnO2 and Fe2O3. Emerson, Carnston & Liss (1979) suggest that the manganese oxide particulate layer found at the oxygen-hydrogen sulphide interface in an anoxic basin (Saanich Inlet, Canada) is a place where there is active copper adsorption. Diagenetic changes within the bottom sediments have been suggested as responsible for maintaining high levels of mercury, cadmium, copper, zinc, and molybdenum in the surface sediments of the upper part of a river estuary (Jones & Jordan, 1979). Palmer & Baker (1978) found copper porphyrins in deep-sea sediments which they suggest are a result of the oxidation of terrestrially derived organic matter

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which has undergone oxidation before or during deposition. Mooney, Bebula, Ferguson & Hallberg (1978), in an attempt to understand the biological and abiological processes leading to formation of stratabound orebodies and to recognition of the environment of deposition, developed a mathematical model to examine the supply of metal species, generation of hydrogen sulphide and chelating agents in an organic-rich zone by bacterial action, diffusion of chemical species through the sediment, formation of metal sulphides and metal complexes, and the competition between these last two processes. METAL—METAL INTERACTIONS HAVING BIOLOGICAL IMPLICATIONS The interaction of metals, including copper, may have an effect on the availability and/or toxicity of an individual metal to marine organisms (e.g., Anderson & Weber, 1976). Kirchgessner, Schwarz, Grassmann & Steinhart (1979) examined the interactions of copper with iron, zinc, molybdenum-sulphur, manganese, nickel and selenium, cadmium, silver, and mercury with respect to agricultural studies. They comment in their introduction that “direct interactions of copper with other trace elements occur when copper is displaced from its complexes by other trace elements or when copper displaces other trace elements from their complexes. These interactions may be explained by the thermodynamic and kinetic stabilities of the particular complexes.” Bliss (1939) defined three different types of joint action of toxicants, namely independent joint action, similar joint action, and synergistic action. The first two may be predicted directly from the known dosage-response curve of the constituents applied alone while, in the third case, the effectiveness of the mixture cannot be assessed from that of the individual components (Braek et al., 1976). One constituent synergizes or antagonizes the other. All three types of joint action have been detected in experiments with freshwater algae. Bartlett, Rabe & Funk (1974) claimed that combinations of copper, zinc, and cadmium were similar in toxicity to equal concentrations of zinc when applied to cultures of Selenastrum capricornutum. Hutchinson (1973) found that copper and nickel acted synergistically and that selenium and cadmium showed antagonism to four other free algae in fresh water. MacInnes (1981) found highly significant toxic synergism in the response of embryos of the American oyster Crassostrea virginica in a 3×3×3 factorial experiment with copper, mercury and zinc as the nitrates and chlorides. The mechanisms of metal synergism are not altogether clear but may be related to the interdependency of uptake and elimination rates of elements relative to environmental levels of biologically available metals (Cross et al., 1973). This can be more specifically associated with either the toxicity of these metals for cells or the presence of common binding sites for which the metals must compete (Khovrychev, 1973). The preferential complexation of metals by natural organics is documented (Schubert, 1954; Goldberg, 1957; Johnson, 1964) and discussed in more detail in the other parts of this section. Literature concerned with the effect of interaction of specific metals with copper is discussed below. Zinc

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In the marine environment, Braek et al. (1976) showed copper and zinc to have a synergistic effect on the phytoplankton species Skeletonema costatum, Thalassiosira pseudonana, and Amphidinium carteri while having an antagonistic effect on Phaeodactylum tricornutum. Gächter (1976) noted that photosynthesis by phytoplankton was significantly reduced due to the synergistic effects of combined metals, including zinc. Eisler & Gardner (1973) found that mixtures of Cu2+ and Zn2+ produced higher mortality of a fish, Fundulus heteroclitus, than expected on the basis of the toxicities of individual components. Nickel In tests with isolates of deep-sea bacteria, Yang & Ehrlich (1976) found growth to be reduced more under the combined effects of copper and nickel than with either of the metals alone. Mercury Khovrychev (1973) found the absorption of copper ions by the yeast Candida utilis to be inhibited in the presence of mercury salts. Photosynthesis in phytoplankton was significantly reduced due to synergism by metals, including mercury (Gächter, 1976) and the bioassays of Gauthier, Bernard & Aubert (1976) showed additions of mercury to increase the toxic effect of copper on the oxygen uptake and growth of the bacterium Staphylococcus epidermidis. Lead Lead salts inhibited the absorption of the copper ions by the yeast Candida utilis (Khovrychev, 1973). The bioassays of Gächter (1976) also included lead which was found to reduce significantly phytoplankton photosynthesis by synergism with copper. Cadmium Combinations of cadmium and copper resulted in a lower rate of oxygen consumption than either metal alone in experiments with Nassarius obsoletus (MacInnes & Thurberg, 1973). The rate of oxygen uptake of distressed and retracted snails was lower than that of controls after exposure to individual metals except cadmium, additions of which resulted in elevation of oxygen consumption. Coombs (1976) showed that the exposure of plaice to sublethal concentrations of cadmium caused significant changes in zinc and copper distributions in tissues. It is believed that these changes could lead to conditional metal deficiencies in blood cells, gills, and serum while mobilizing liver and kidney stores with the production of low molecular-weight proteins suggestive of metallothioneins. Cadmium, in combination with copper, was also found to reduce significantly photosynthesis in marine phytoplankton (Gächter, 1976) and to have no modifying effect on the toxicity of the metals on the bacterium Staphylococcus epidermidis (Gauthier, Bernard & Aubert, 1976). Eisler & Gardner (1973) noted that concentrations of cadmium

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not normally lethal became lethal to fish exposed to copper and copper plus zinc. Iron Luoma & Bryan (1978) found the availability of copper to two species of deposit-feeding clams to be inversely related to the iron content of sediments. Manganese The experiments of Yang & Ehrlich (1976) on deep-sea bacteria isolates suggested that growth was depressed more by a combination of manganese and copper than by either of the metals alone. Sunda, Barber & Huntsman (1981) note a physiological interaction between copper and manganese in phytoplankton, in which copper competes for manganese nutritional sites thereby interfering with manganese metabolism. They indicate that the ratio between the cupric and manganese ions should change in upwelled water as it ages at the surface, producing a change in the effect of copper on manganese uptake and, thus, on cell division. Magnesium Yang & Ehrlich (1976) found that magnesium reduced the toxicity of copper, nickel, and cobalt when combined. Cobalt The growth experienced by the deep-sea bacteria cultures of Yang & Ehrlich (1976) when subjected to copper and cobalt in combination, was less than that realized by cobalt alone but greater than that of copper alone. Because of the interdependency of metals in biological systems, Coombs (1976) expressed strong opposition to single element analyses and felt that a multi-element investigation is the only realistic approach in pollution studies.

MANGANESE NODULES, MINING, AND DREDGING This section deals with the use of copper-containing antifouling compounds and what could be classified as mining-type industrial operations, including dredging. It includes the effects that they have on copper levels both in the environment and in organisms. The deposition of metals in the oceans can occur through both diagenic and biogenic processes. Armstrong & Miall (1945) comment that “what the chemist does with difficulty the oyster finds easy. Apparently it gargles a barrel of water per day and around the British Isles and in certain parts of the Atlantic coast oysters become green due to the formation of a respiratory pigment, haemocyanin, containing copper.” ANTIFOULING COMPOUNDS

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Copper has proved extremely useful in the disinfection of water (e.g., Muzychuk, 1977), as an antibacterial agent (e.g., Foye, 1977), as an algicide (e.g., Dempster & Shipman, 1969), as a pesticide (e.g., Zdybiewska & Kluczycka, 1974; Roegge, Rutledge & Guest, 1977; Cheng, 1979), and as an antifoulant (e.g., Clapp Laboratories, 1964; Southwell & Bultman, 1970; Norman & Henningsson, 1975; Tamblyn, Rayner & Levy, 1978; Marshall, 1979; Tamblyn & Levy, 1979), as well as a means of preventing intrusion by marine organisms (e.g., Muraoka, 1970) as in the use of copper strips to prevent oyster predation (Huguenin, 1977). Exposing structural surfaces to marine environments or using salt water as a coolant in tubing allows the growth of marine organisms (biofouling) with the subsequent increase in weight and friction against water movement. This leads to premature failure of structures such as pilings, to reduced speeds and greater power requirements in ships, and to reduced flow in salt-water cooling systems with resultant decrease in efficiency. Biofouling is not a simple process as there are numerous events that occur in the succession of organisms on a new surface (e.g., Gerchakov & Sallman, 1977). The production of antifouling compounds is a multi-million dollar industry and as MacArthur (1968) points out “the requirements for prolonged submersion imposes special problems of corrosion, marine fouling, and biodeterioration”, a situation which requires suitable antifouling materials (e.g., Phillip, 1973a). These include copper-containing paints and coatings as well as controlled leaching of copper surfaces by electrolysis (Spears, Stone & Klein, 1969). With copper-containing compounds, a suitable material means one in which enough copper is available to either prevent the settlement of fouling organisms or to kill them after they have settled. The amount of copper necessary to do this varies; Groover et al. (1970) suggest that to remain free from fouling organisms a minimum rate of 5 mg·dm−2·day−1 of copper must be leached from the surface while De Wolf & Van Londen (1966) indicate that an antifouling compound containing copper as the toxicant will prevent fouling as long as the leaching rate is >10 µg·cm−2·day−1 (=1 mg·dm−2·day−1). Phillip (1973a) also found this amount suitable to prevent settlement of most organisms. De Wolf (1964) and De Wolf & Van Londen (1966) point out that variation in copper release over a surface can explain the irregular distribution of barnacles although certain organisms exhibit the ability to adapt to copper (Starostin, 1968). Groves, Dennon & Peterson (1970) discuss corrosion properties and potential value for the use of copper alloys with cathodic protection. Ritter & Suitor (1976), in a comparison of heat exchange tubing, found that a copper-containing alloy (Alloy 706) did not harbour marine organisms as much as a titanium-containing tubing when exposed to silt fouling at low surface temperatures. Cologer & Freiberger (1967) utilized a technique for rearing and selecting standardized populations of young attached barnacles for the evaluation of experimental antifouling paints. Three species of barnacles were used as earlier work (Weiss, 1948) had shown that the tolerance to copper in barnacles varied from species to species. Weiss (1947) had also found that tolerance varied from one group of fouling organisms to another. Phillip (1973a) discusses the development of coatings to overcome the problems of varying environmental conditions that affect the efficiency of copper-containing antifouling compounds. He, as well as others, (e.g., Castelli, 1977; Marshall, 1979), however, emphasize the need to develop antifouling coatings to replace copper oxide as

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the main toxic component and considers organometallic compounds without copper but including alkyl-mercury compounds. Garey (1979) describes the use of organometallic coatings as well as cuprous oxide on condenser tubing to reduce slime while O’Neill & Mathews (1979) mention the incorporation of cuprous oxide into marine concrete to reduce fouling. The development of suitable copper-containing alloys (e.g., Ritter & Suitor, 1976) also is an area of major interest, to produce an antifouling substratum with a minimum of peripheral environmental effect. The evaluation of ‘Ocean Thermal Energy Conversion’ as a source of power has caused the examination of material useful in the prevention of fouling (e.g., O’Neill & Mathews, 1979) as well as a compilation of existing data regarding selected sea-water properties thought to affect biofouling and corrosion (Craig et al., 1978). The importance of certain environmental conditions, on the chemistry of copper-containing surfaces is discussed by Ikemura (1969) as are the effects of biofouling on corrosion. Phillip (1973b) suggests that fundamental studies need to be done on membrane properties, enzymes, and adhesives of fouling organisms to reduce the use of toxic substances. There is also a need for application of existing ecological information to biofouling. Hanson & Bell (1976) for example, found that fouling decreased with depth and suggested that the cooling system of a proposed power plant should be in water deeper than 6 m to reduce fouling. The use of copper in biocidal agents has been of concern not only because of the potentially toxic effects on other organisms, but also on uptake by economically important species. The estimated total release of copper from antifouling coatings into the marine environment is appreciable (e.g., Phillip, 1973b). D.R.Young et al. (1979) point out that the estimated application rate of copper via copper bottom paints in southern California harbours alone, is 180 metric tons·yr−1, more than one-third the annual emission rate from major municipal waste waters (510 metric tons·yr−1), and twice the estimated input from surface run-off and dry aerial deposition (≈70 metric tons·yr−1). The release of this material means that it may be available for uptake by organisms. Although Young et al. (1974) imply that sediments near boat marinas may accumulate copper, this appears to depend on water circulation as well as chemical conditions in the water and sediments (de Mora, Wong & MacDonald, 1978; Compton & Corcoran, 1975). DEEP-SEA METAL DEPOSITS Humphris & Thompson (1978) examined trace element mobility during hydrothermal alteration of oceanic basalts and found that copper was leached from the basalt but often precipitated as a sulphide, forming vein-like localized deposits. McArthur & Elderfield (1977) discuss the metal accumulation rates in sediments near a mid-oceanic ridge in the Indian Ocean, suggesting that metalliferous deposits are probably local and not widespread. The potential for various methods of mining oceanic deposits of metals as well as the potential of the deposits and of the environmental problems associated with exploration and mining has been examined by a United Nations Joint Group of Experts on the Scientific Aspects of Marine Pollution (GESAMP, 1977). Glasby (1979) provides a general discussion on the sources of oceanic minerals, including brine pools, manganese nodules, and metalliferous sediments. Of these three the current favourite is manganese nodules with a great deal of information becoming available on the nature and

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formation of the nodules, their association with the adjacent environment, the problems and effects of recovery, and the economic aspects of the developing manganese nodule industry. Certain geographic areas may form important geochemical environments for the accumulation of trace elements (Oldnall, 1975; Murty, Rao, Veerayya & Reddy, 1977). Although native copper has been found only once in sediments (Siesser, 1976) copper is common in other forms (e.g., Brockamp, Goulart, Harder & Heydermann, 1978) and has been found in bedrock, in geothermal deposits, and in ferromanganese nodules. The association of copper with hydrous oxides, through processes of adsorption and coprecipitation, make ferromanganese nodules an important source of the metal (e.g., Ehrlich, Yang & Main waring, 1971; Arrhenius, 1974; Bender, 1975; Morgan & Moore, 1975; Uchio, 1979). The deposition of metals in the formation of nodules, or encrusting ferromanganese deposits is due in part to microbial activity (McLerran & Holmes, 1974; Ehrlich, 1975) and possibly the associated activity of other organisms (Dugolinsky, Margolis & Dudley, 1977). Ehrlich (1975) suggested that Mn(II)-oxidizing bacteria can assist in the growth of manganese nodules although the nature or state of the metals in the sediments exerts some control in nodule formation (Bowser & Mills, 1978). Frazer et al. (1978) discuss the chemical availability of copper, nickel, cobalt, and manganese for ferromanganese nodules while Fanger, Pepelnik & Grondey (1977) discuss a neutron activation method of analysis for metals in manganese nodules. Pettis & De Forest (1979) analysed ferromanganese nodules from the Southern Ocean, south-west of Cape Leeuwin (Australia) for a number of elements and made a preliminary attempt to compare the results of the analyses with those of nodules from areas in the South Pacific, Atlantic, and Indian Oceans. Halbach, Rehm & Marchig (1979) examined the distribution of a number of metals, including copper, in grain-size fractions of sediment from a manganese nodule field in the central Pacific Ocean. They found that more than half of the copper was bound in the <2 µm-fraction and that there was a good positive correlation between the concentrations of manganese, nickel, copper, and cobalt which they discuss in terms of macronodule growth. Rutkovskiy (1977) found that constants for calcium, magnesium, zinc, cobalt, copper, and nickel exchange for sodium in ferromanganese nodules from the Pacific Ocean indicated a high capacity of the nodules for adsorbing certain ions, a condition which would indicate a mechanism for trace metal accumulation by the nodules. Usui (1979) on the other hand examined the 10-A manganite formation in the laboratory and found that there is a rapid transformation through interactions with ions, their incorporation and the release of Na+, to form the 10-A manganite. From this he concluded that the reacting ions (Ca, Mg, Ni, Cu, Co, and Zn) are apparently located in the definite cation sites of the crystal structure of 10-A manganite in deep-sea nodules. Glasby, Keays & Rankin (1978) measured a large number of metals in deep-sea manganese nodules and coexisting sediments from two nodule sites in the western Pacific basin as well as in one fossil nodule from Timor. Copper in the deep-sea nodules ranged from 0.15–0.37% and was 0.28% in the fossil nodule. Sedimentological and structural investigations of ferromanganese accretions or nodules suggests that they most frequently form in waters deeper than 200 m (Xavier, 1976), although Djafari (1977) and others have found manganese-iron accumulations in shallow water. Crerar & Barnes (1974) and Horn, Horn & Delach (1972) discussed conditions

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under which manganese nodule formation might occur and changes in composition as a result of the nature of the sediment. Horn et al. (1972) pointed out that the nickel and copper levels of nodules found in siliceous oozes are approximately twice as high as the levels on nodules found in red clays. McKenzie (1975) compared the relationships between manganese and several other metals in both terrestrial soils and marine manganese nodules while Sorem, Reinhart & Fewkes (1977) analysed 140 manganese nodules for concentrations of Mn, Fe, Ni, Cu, Co, and Zn. Jenkins & Lense (1967) explored the area of Norton Sound, Alaska, to evaluate the mineral potential of the area. Ottow (1978) examined mechanisms of ferromanganese nodule formation in the sediment-water interface of the deep sea and under the mechanism of accretion by enzymatic oxidation of adsorbed Mn(II) ions suggested first a non-biological adsorption process and then subsequent enzymatic oxidation by Mn(II)-oxidizing bacteria. It is suggested that in this way, new adsorption sites for further reactions are produced, sites with affinities for iron, copper, nickel, and cobalt. Schnier, Gundlach & Marchig (1978) suggested that copper in the pore water and sea water in the radiolarian ooze area of the central Pacific may be mobilized by the dissolution of radiolarian tests as well as by the dissolution of manganese micro-nodules in the sediment. The importance of biological processes in the formation of nodules is suggested by Ghiorse (1979) in the abstract of a paper on the potential of hypho-microbia to be important metal depositors on the surfaces of ferromanganese concretions; this is also suggested by Thiel (1978). The importance of bacteria to the formation of manganese nodules (Ehrlich, 1975) combined with the variation in metal composition of the nodules produced by the types of sediments in which the nodules are found (Horn et al., 1972) suggests an interaction of the metals, the environment, and the micro-organisms. Yang & Ehrlich (1976) and Arcuri & Ehrlich (1977) for example, found a variation in the effect of copper on the growth of three deep-sea bacterial isolates which could be partly associated with changes in pressure. One of the potential problems of manganese nodule dredging is that plaguing any dredging operation, the potential chemical changes in the water column in the area of the operation and the biological effect of those changes. Bischoff & Rosenbauer (1978) examined chemical changes in sea water caused by resuspension of deep-sea sediments from DOMES sites A, B, and C and found only slight changes in all components except silicate, in both oxic and anoxic situations. Deep-ocean mining has been contemplated for a number of years, ever since the 1873 Challenger Expedition recovery of metalliferous nodules (Mero, 1978). The discovery of metalliferous muds in the Red Sea in the early 1960s (e.g., Mero, 1978) and the finding of massive deep-sea sulphide ore deposits on the East Pacific Rise, with at least 29% zinc metal and 6% metallic copper (Francheteau et al., 1979) provides an adequate supply of ore. Varentsov (1979) found localizations of copper and other metals in upper Cenozoic sediments near the crest of the Mid-Atlantic ridge, suggesting some hydrothermal concentrating mechanism. Mustaffi & Amann (1978) discuss the nature of Red Sea metalliferous muds, the problems of mining them, and environmental considerations associated with exploration and mining. The question as Halkyard (1979) points out is when deep-ocean mining will be economically and technically feasible. Mining localized deposits in estuaries, originating from fluvial sands and anthropogenic input has also

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been considered but Lyons & Gaudette (1979a) suggest that, with today’s technological standards, these accumulations are about an order of magnitude below economically minable ores although if necessary, technology could be developed for future mining of these deposits. Deep-sea mining involves a variety of environmental problems (e.g., Portmann, 1971; Padan, 1971), most of which would have little effect on coastal fisheries. Mines in coastal regions, however, frequently discharge tailings and overburden into the marine environment. In a sense, the discharge of tailings can be compared to a dredge-spoil site in which a specific type of material is discharged over an extended period. Mamen (1973), for example, discusses the nature of a copper ore body in an open pit mine on Vancouver Island, Canada, with tailings which are discharged into the marine environment. He provides details of the metal extraction process and generalities of the ecological monitoring programme. Castilla & Nealler (1978) discuss the environmental impact of mining activities of a copper mine in Chile. They conclude that “chemical pollution and sediment accumulation hinder the development of benthic invertebrates, algae, and fish”. Thompson & McComas (1974) determined copper and zinc levels in submerged mine tailings at Britannia Beach, British Columbia, Canada, and found concentrations of copper ranging from 83–1394 µg·g−1 sediments. Thompson & Paton (1978), in the same area, found that copper in the sediment interstitial waters was higher than that in the overlying water and related this to the higher levels of copper in the mine tailings. Marvin, Lansford & Wheeler (1961) attempted to control a red tide organism (Gymnodinium breve) by introducing 60 tons of a raw sulphide containing 1% copper ore into a lagoon with a volume of 2.84×105 m3. The results indicated that the copper concentration in the lagoon was not thereby increased to a level lethal to G. breve and that the flora and fauna of the lagoon showed no significant effects which could be attributed to copper. DREDGING Dredging of benthic sediments for navigational purposes and as a source of sand, gravel, and shell for construction purposes has been considered as a potential environmental hazard (e.g., Ecker & Hendricks, 1976). The problems of dredging include siltation and re-mobilization of heavy metals from the sediment. The concerns about the potential hazards of dredging have stimulated an examination of dredge sites as well as dredgespoil dump sites to determine conditions before (e.g., Soule & Oguri, 1976) and after dredging (e.g., Anderlini et al., 1975; Serne & Mercer, 1975) or dredge-spoil dumping (e.g., Palmer & Lear, 1973). Whitfield (1976) described the shell dredging operation in Florida and the involvement of governmental agencies. Anderlini et al. (1975) examined the distribution of metals in dredge-spoil areas and found that most of the copper in the interstitial water was associated with soluble organics with a molecular weight of <10 000. Stern (1975) found that copper was adsorbed almost immediately by a portion (<62.5 µm) of a sediment sample used to simulate dredge-spoils. Ali, Gross & Kishpaugh (1975) found that sites containing dredge-wastes from New York Harbour could be identified on the basis of their chemical properties suggesting retention of unique properties after dumping. Trefry, Sims & Presley (1976) and Trefry & Presley (1976)

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noted that, with shell dredging, there was no indication of an increase in the dissolved metal content in San Antonio Bay, Texas. The ecological effects of dredging and dredge-spoil disposal are reviewed by Morton (1977) and Bouma, Hall & Sidner (1976). Morton (1977) discusses the effects of changes in circulation patterns, re-distribution of sediments, chemistry of sediments, remobilization of toxic materials, topographic features, burial of organisms, and bioconcentration of materials. There is a decrease in species diversity as well as a decrease in the abundance of species that are present in the region of a dredge-site as well as near dredge-spoil disposal sites (Rosenberg, 1977; National Marine Fisheries Service, 1976a,b) although in the latter it is limited to the immediate area of spoil-disposal. Patrick et al. (1977) discuss the physico-chemical factors regulating solubility and bioavailability of toxic heavy metals in contaminated dredged sediment. Studies of the potential biological effect of dredging in San Pedro Bay, California (Soule & Oguri, 1976) suggest that dredging might reduce survival of certain organisms although the nature of the limiting factors was not clearly identified. The results of studies in two widely separated areas (Anderlini et al., 1975; Trefry & Presley, 1976) suggest that the impact caused by dredging is short-lived and that changes in metal concentrations in sediment and invertebrates are small. This does not, however, apply to continuously used dredge-spoil sites or to continuous dredging in an area where increases in turbidity can have an effect on economically important organisms such as oysters. In a paper presented at the ninth dredging seminar Slowey & Neff (in Herbich, 1977) state that only metals soluble within the interstitial water and readily released to the water column are available to a clam (Rangia cuneata) and a shrimp (Palaeomonetes sp.). They observed little uptake with true deposit-feeding infauna and suggest that sediments such as dredged material, with high metal loadings do not necessarily have an abundance of biologically available metals. They also suggest that bulk metal analyses are not useful criteria for evaluating the environmental impact of dredged material on benthic organisms. Smith (in Herbich, 1977), at the same seminar, reported that dredged material offers a medium suitable for marsh, terrestrial, island, and aquatic habitat development. Hirsch et al. (1978) report that biological uptake of sediment-associated heavy metals from dredged material is the exception rather than the rule and, like Slowey & Neff, indicate that bulk sediment content of heavy metals is not correlated with environmental impact. In an attempt to determine if dredging operations in Mare Island Strait, in the San Francisco Bay region released heavy metals, including copper, Anderlini et al. (1975) measured metal levels in both sediments and invertebrates and found that changes in mean metal concentrations in sediments and invertebrates during the dredging operation were small (
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following the disposal of polluted sediments. This latter may very well have been pore water organometallics released upon disturbance of the sediments. Sustar & Wakeman (1976) found that the release of copper from sediments under oxidizing conditions increased with agitation suggesting that a kinetic mechanism may play a rôle in the fate of trace metals. They did not, however, find an obvious biological uptake of metals as a result of dredging operations. Swartz, Deben & Cole (1979) developed a bioassay technique for the toxicity of dredged sediments to marine macrobenthos, using five macrobenthic invertebrates which represented different taxonomic and trophic positions. The apparatus used in the experiments, however, showed little, if any, sensitivity to burial under 15 mm of sediment or to alterations in the distribution of particle size. The bioassay was not designed to evaluate anything except the physical process of sediment burial and the authors comment on the need for the effects of other factors to be determined. The bioassay test for contaminant uptake from dredged material by marsh plants, developed by Wolf et al. (1978) suggests that there is species specificity in copper accumulation as well as some indication of differences from one type of dredged material to another. Windom (1977) examined the ability of salt marshes to remove heavy metals from dredged materials and found that copper accumulated in the sediments more than in the marsh grass (Spartina alterniflora) roots or leaves. Keillor & Ragotzkie (1976) found no evidence of bio-magnification of copper, chromium, arsenic. cobalt or selenium in the food chain of Lake Superior as a result of disposal of dredged material, even although copper and arsenic were at concentrations more than twice the background values in some sediments. Part of the reason for the apparent low accumulation of copper is that much of the metal is bound in mineral lattice sites and is essentially inert (e.g., Serne, 1977) while a second reason is that the relatively high organic content of many sediments includes high levels of complexing agents which reduce the biological availability of the metal (e.g., Lewis et al., 1973).

COPPER LEVELS IN THE MARINE ENVIRONMENT Krauskopf (1956) estimated 42 mg Cu·1−1 to have been added to the world’s oceans over geologic time. Present levels are 4 or 5 orders of magnitude below this as a result of complex chemical relationships involving solubilities, saturation levels, redox reactions, precipitation, deposition, biological uptake, and metal absorption/desorption phenomena as well as complexation by organic and inorganic components of sea water. With respect to CuCO3, the least soluble compound which copper forms with one of the anionic species in sea water, copper appears to be under-saturated in sea water by 150–300 fold. The average concentration of copper in sea water generally reported in literature prior to the mid 1970s is about 3 µg·1−1 (Goldberg, 1965). Much lower levels have been reported in more recent literature (e.g., Boyle & Edmond 1975) (see Table I). To facilitate the organization of the data, Tables II through VI represent copper levels at the air-sea interface, “dissolved” in sea water, in particulates, in interstitial water, and in sediments. The data must be considered with caution, especially when comparing the results of the different studies which are reviewed in this report, as the number of techniques and methods of presentation of the data is astounding. The reader will

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encounter wet weights, dry weights, ash weights, percentage weights, total, dissolved, and particulate copper values, copper values for varying size fractions and numerous other variations on the above themes. We have not attempted to present all of the data encountered in reviewing the literature and numerous values have been omitted for reasons of clarity, brevity, repetition, or oversight. A number of reviews of environmental copper levels exist; two of the most recent and extensive are those of Schmidt (1978a, b) who has summarized current and historic literature on levels in the marine environment and in marine organisms and provides a detailed discussion of the biogeochemistry of copper in the sea. Jørgensen (1977) has prepared the most recent and extensive review of values occurring in the literature. Other reviews include Bouquiaux (1974) on metal levels in organisms and the environment, Burton (1978a) with a general review of metal speciation in estuaries, Förstner et al. (1978) with a discussion of sources of metals in sediments, Jenne & Luoma (1977) on the bioavailability and metal forms, uptake and release mechanisms, Leland et al. (1979) on bioaccumulation and toxicity, Bascom (1978) with a compilation of the constituents of major effluent and wastewater discharges, and Bernhard (1978) on heavy metals in the Mediterranean.

TABLE I Example of fluxes of copper to the marine environment: units unless stated otherwise are kg/yr; P=paniculate; D=dissolved. *from review paper, not from original

Location Asia

Level

Comment

Reference

D

P

3.53 (×103 ton/yr)

1.04 (×103 ton/yr)

1.26 ”

0.036 ”

”

0.634 ”

1.24 ”

”

1.25 ”

7.23 ”

”

7.8 ”

11.9 ”

”

Kara Sea

14.4 ”

9.88 ”

”

Laptev Sea

6.86 ”

2.76 ”

”

Pacific (Bering, Okhotsk, Japan seas)

12.1 ”

4.45 ”

”

Copper discharge from U.S.S.R. into: Barents and White Seas Baltic Sea Black and Azov Seas (without Danube) Danube River Aral-Caspian Basin

Atlantic South Eastern Atlantic Saltmarsh Estuaries:

Konavolov & Ivanova, 1971

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D

P

75×103

5×103

Windom, 1975

Black River

5×103

1×103

”

Santee River

5×103

5×103

”

Cooper River

42×103

11×103

”

Savannah River

17×103

32×103

”

Ogeechee River

8×103

2×103

”

Altamaha River

41×103

19×103

”

Satilla River

11×103

2×103

”

18×103

5×103

”

Pee Dee River

St Johns River Europe-Atlantic Coast Rhine River discharge (fraction<16 µm)

1100 ton/yr

1775 ton/yr

DeGroot, 1973

D+P Rhine River discharge Netherlands (other rivers 5% Rhine)

Location

2900×106

Level

Hueck, 1975

Comment

Reference

North America–Atlantic Coast Fluxes to Narragansett Bay sediments:

Fluxes to Chesapeake Bay sediments:

193µg·cm−2·yr−1

Anthropogenic

Goldberg et al., 1978a*

3.1 µg·cm−2·yr−1

Natural

”

4.5 µg·cm−2·yr−1

Core 1411, anthropogenic

”

4.0 µg·cm−2·yr−1

Natural

”

0.8

µg·cm−2·yr−1

Core 1314, anthropogenic

”

3.6

µg·cm−2·yr−1

Natural

”

Input at dumpsites: Du Pont

2400

Philadelphia

32.6×103

Upper Newport Bay, 52 µg·cm−2·yr−1 Maryland

Pesch et al., 1977

Total input

Christensen & Scherfig, 1978

Interlinking of physical

Key West, Florida

22.7–45.4 kg/day Desalination plant discharge 143 µg/1 (max. 6515)

Chesapeake Bay, Maryland

New York City

1112

Chesher, 1971

Average effluent

50–2000×106 g/yr Wastewater discharge (concentration Schmidt, >50 µg/1) 1978a* 200×106 g/yr

Mean fluvial discharge (concentration in major source of freshwater to Chesapeake Bay=4.2 µg/1)

900×106 g/yr

Discharge to sea

”

Long Island Sound, Conneticut-input by: Connecticut River

6×107 g/yr

Dehlinger et al., 1974

Thames River

0.8×107 g/yr

”

Housatonic River

13.4×107g/yr

”

Other surface flows

3.3×107

g/yr

”

Effluents: primary

8.7×107

g/yr

”

secondary

2.1×107g/yr

”

North America—Pacific Coast Puget Sound—Copper inputs: Rivers

787×103

Schelle & Nevissi, 1977

Metro, West Point Plant

29×103

”

Other Municipalities 22×103

”

360–590×103

Vessels

Location

”

Level

Comment

Reference

15×103

”

Advective transport

306×103

,,

Atmospheric input

450×103

„

Seattle, urban run-off

Los Angeles, natural copper input from: S.California Rivers -sewage input -natural copper levels -waste water

3 ≈3×10 3 ≈420×10 1–10 µg/1

100–1100 µg/1

Galloway, 1972a*

Oceanography and marine biology

639

Southern California: The Bight (Santa Barbara to San Diego) 570×103

-wastewaters

Young et al., 1974

19×103

-run-off -vessel antifouling paints Fluxes to Santa Barbara Basin sediments:

180×103 1.4 Anthropogenic µg/cm2/yr

Goldberg et al., 1978a*

2.6 Natural µg/cm2/yr San Francisco Bay Estuary inputs

200×103 Suspended river load

Eaton, 1979b

100×103 Dissolved river load 16×103 Ocean 10×103 Aerosol 2×103 Dissolved storm run-off 5×103 Total storm run-off 20×103 Dissolved sewage 180×103 Oxidizable and desorbable sewage 3×103 Dredging release Discharge from 8 Edison Generating plants

2.1×103

Young et al., 1977

TABLE II Copper levels at air-sea interface (air, rain, aerosols)

Location

Level

Reference

a) AIR Atlantic Bermuda 1973

≤0·08–24 ng/Standard m3 (mean=0·90;

1974

60 samples)

0·07–15 ng/Standard m3 (mean=1·2; 74 samples)

Coastal Australia/New Zealand

Duce et al., 1974

Interlinking of physical

Hobart and Derwent Estuary, Tasmania

1114

311–4400 µg/g (dust settled in 1 month, Bloom & Ayling, 18 stations) 1977

Coastal Great Britain Plynlimon, Montgoms

11 ng/kg air

Peirson et al., 1974

Styrrup, Notts

34 ng/kg air

”

Trebanos, Glamorgs

29 ng/kg air

”

Waymires, Lancs

<6 ng/kg air

”

Chilton, Berks

13 ng/kg air

”

Gresham, Norfolk

<9 ng/kg air

”

Arran, Bute, Scotland

<5 ng/kg air

”

Leiston, Suffolk

<11 ng/kg air

”

Collafirth, Shetland Is

<4 ng/kg air

”

Lerwick, Shetland Is

<8 ng/kg air

”

<6 ng/kg air

”

<17 ng/kg air

”

Arran, Bute, Scotland

27 µg/l

”

Leiston, Suffolk

28 µg/l

”

Gresham, Norfolk

27 µg/l

”

Collafirth, Shetland Is

23 µg/l

”

Lerwick, Shetland Is

23 µg/l

”

41 µg/l

”

0·5–60·0 µg/l (generally not >10·0)

Biggs & Miller, 1973

82 µg/l

Peirson et al., 1974

0·12–10 ng/m3

Cattell & Scott, 1978*

Europe Atlantic Coast Petten, N.Holland North Sea Gas Platform b) RAIN Coastal Great Britain

Europe Atlantic Coast Petten, N.Holland North America—Atlantic Coast Delaware North Sea Gas Platform c) AEROSOLS Atlantic North of 30° N

Oceanography and marine biology

641

Antarctic 0·025–0·064 ng/m3

”

0·3–190 ng/m3

”

Ensenada, Mexico

20·0 ng/m3

Hodge et al., 1978

La Jolla, California

16·0 ng/m3

”

San Francisco, California

34·0 ng/m3

”

South Pole Coastal Australia/New Zealand Tasmania North America—Pacific Coast

COPPER LEVELS IN MARINE WATERS Copper levels encountered in nearshore waters are generally higher than those given for open-ocean samples due to the proximity of copper sources. Schmidt (1978a,b) found the mean copper concentration quoted in the literature for coastal waters to be ≈2.0 µg·1−1 (excluding inordinately high values) and calculated a nearshore to open-ocean mean copper concentration ratio of about 1.7:1. He notes copper levels in surface waters ranging from 0.06–6.7 µg·1−1. Eaton (1979a) observed copper concentrations in the offshore waters of the San Francisco Bay Estuary to be extremely low and believes variations in soluble trace metals in the Estuary to be governed predominantly by physical mixing processes. He also notes an excess of copper in the lower Estuary which he ascribes to anthropogenic influences. No systematic onshore-offshore trends in copper concentrations were found in surface waters off the southern Atlantic coast of the U.S.A. by Windom & Smith (1979) while regional variations were found to be related to the source of the water. Copper concentrations tend to vary inversely with depth (Schmidt,

TABLE III Copper levels in sea water: unless otherwise stated, levels are for filtered water (any filter type) and are in µg/l; (T), soluble+paniculate copper. *a review paper containing levels used in determining the concentration range; this table is for ordering selected references rather than providing detailed levels of copper

Location

Range

References

World Oceans Open ocean

0·1–3·9 (T)

Average

0·8

Nearshore

0·3–3·8 (T)

Chester & Stoner, 1974; Phillips, 1977*

Chester & Stoner, 1974; Phillips, 1977*

Interlinking of physical

Average

1116

0·9

Atlantic Ocean and adjacent seas Eastern Atlantic

0·069– 24·8

Meulen, 1931; Bardet et al., 1938; Frache et al., 1976*; Moore & Burton, 1976; Moore, 1978

1·9– Phillips, 1977* 107·0 (T) Northeastern Atlantic

Barents and Norwegian Seas

0·05–88

Barnes & Rothschild, 1950; Chester & Stoner, 1974; Frache et al., 1976*; Jørgensen, 1977*; Romeril, 1977

0·05– 0·80 (T)

Phillips, 1977*

1–23

Petrov et al., 1976*; Melhuus et al., 1978

3·7–77·0 Phillips, 1977* (T) North Sea

0·1–34

Dutton et al., 1973; Hueck, 1975; Thornton et al., 1975; Petrov et al., 1976*; Duinker & Kramer, 1977; Jørgensen, 1977*

2·8 (T)

Phillips, 1977*

Dutch Waddensea 0·1–>30 (T)

Jørgensen, 1977*

Baltic Sea

Noddack & Noddack, 1939; Buch, 1944; Kremling, 1973; Petrov et al., 1976; Jørgensen, 1977*; Weigel, 1977; Magnussen & Westerlund, 1980

0–22·8

Location

Range

References

English Channel

0– <3000

Atkins, 1932, 1933, 1953; Black & Mitchell, 1952; Thornton et al., 1975; Frache et al., 1976*; Jørgensen, 1977*; Romeril, 1977

Irish Sea

0·18– 12·5

Elderfield et al., 1971; Halcrow et al., 1973; Thornton et al., 1975; Foster, 1976; Petrov et al., 1976*; Jørgensen, 1977*; McGrath & Austin, 1979

0·5– 17·3 (T)

Morris, 1974; Phillips, 1977*

Equatorial Atlantic and Sargasso Sea

0·087– 1·45

Bender & Gagner, 1976; Frache et al., 1976*

Southeastern Atlantic

0·3–17

Orren, 1967; Chester & Stoner, 1974

Northwestern Atlantic

<0·10– Hiltner & Wichman, 1919; Prytherch, 1934; Galtsoff, 1943; 600 Duce et al., 1972; Dehlinger et al., 1974; Arnac, 1976; Bewers et al., 1976; Frache et al., 1976*; Duedall et al.,

Oceanography and marine biology

643

1978; Waldhauer et al., 1978 0·56–16 Piotrowicz et al., 1972; Phillips, 1977*; Yeats et al., 1978 (T) Southwestern Atlantic including the Gulf of Mexico and the Caribbean Sea

0·36–44 Riley, 1937; Galtsoff, 1943; Lowman et al., 1966; Alexander & Corcoran, 1967; Leyden et al., 1975; Frache et al., 1976*; Trefry et al., 1976; Schmidt, 1978a* 1·1–2·4 Montgomery & Santiago, 1978 (T)

Mediterranean and adjacent seas

Indian Ocean and adjacent seas, including the Red Sea

0–2500 Dieulafait, 1879; Fonselius, 1970; Frache et al., 1976*; Fukai & Huynh-Ngoc, 1976; Petrov et al., 1976*; Jørgensen, 1977*; Roth & Hornung, 1977; Barbara et al., 1978; Benon et al., 1978; Shiber, 1979 0.1– 22.4 (T)

Phillips, 1977*

0.080– 76.12

Kappana et al., 1962; Fonselius, 1970; Sankaranarayanan & Reddy, 1973; Chester & Stoner, 1974; Frache et al., 1976*; Danielsson, 1980

1–5 (T) Phillips, 1977* Pacific Australia-New Zealand

0.06–29 Brooks, 1965; Boyle & Edmond, 1975; Bloom & Ayling, 1977; Blutstein & Smith, 1978

Northwestern Pacific and 0.3–60 adjacent seas

Isibasi et al., 1940; Morita, 1950; Chester & Stoner, 1974; Inoue et al., 1974a; Frache et al., 1976*; Kusaka et al., 1977; Wong & Li, 1977; Naruse et al., 1979

2.5–160 Phillips, 1977* (T) Northeastern Pacific

0–29.0

Severy, 1923; Chow & Thompson, 1952; Serne & Mercer, 1975; Burrell, 1976; Ford et al., 1976; Petrov et al., 1976*; Cave, 1977; Bruland, 1978; Franks & Bruland, 1978; Heggie & Burrell, 1978; Sugai & Healy, 1978; Zirino et al., 1978b

0.8–5 (T)

Phillips, 1977*

1978a, b). There does, however, appear to be an increase in copper at about 500–1000 m, possibly related to the presence of sinking, decaying organisms (Riley & Taylor, 1972; Slowey et al., 1967). Forster, Kharkar, Turekian & Thompson (1978) discuss the transfer of materials in the water column and suggest that “the trace elements that are biologically active and associated with suspended material in the upper water column appear to be transported to depths for subsequent uptake by benthic organisms. This is shown by

Interlinking of physical

1118

benthos/plankton ratios which are greater than one for manganese, iron, cobalt, nickel and cadmium. The more soluble species of copper and chromium show benthos/plankton ratios less than those which indicate a rapid recycling of those species in the mixed layer.” Danielsson (1980) also suggests that the distribution of copper in vertical profiles in the Indian Ocean is controlled by biogenic processes. He observed copper levels to be higher at the bottom than at the surface with a minimum at mid-depths. Determinations of Baltic Sea water showed metal concentrations to be generally independant of depth for all metals examined except copper which exhibited a small but significant decrease below 80 m (Magnusson & Westerlund, 1980). Burrell (1973) found total extractable copper values to increase near the bottom of an Alaskan sub-arctic fjord. Work in estuarine waters of the Yarra River, Australia, (Blutstein & Smith, 1978) showed a similar near-bottom increase with a high near-surface value in low salinity water. Copper levels also tend to decline with increased distance from the mouth of a river (Sick, 1978; Duinker & Nolting, 1976). In an examination of very low salinity regions of estuaries Morris et al. (1978) found a mid estuarine maximum corresponding to a natural input of metal-rich solar-warmed pore water from the large areas of intertidal mud flats in the Tamar estuary, U.K. Dissolved and paniculate copper decrease with salinity in the River Rhine Estuary (Duinker & Nolting, 1978). Increased copper values are also associated with nearshore upwelling waters (Knauer & Martin, 1973; Cave, 1977) and zones of relatively strong convergence (Sick, 1978). Seasonal fluctuations of copper occur due to fluctuations in run-off, with increased values in the spring and summer (Chow & Thompson, 1954; Thomas & Grill, 1977; Foster & Morris, 1971; Arnac, 1976) at the time of the heavy spring run-off. Atkins (1953) found decreased values in the English Channel during the same period, due to biological uptake of copper, as did Schubel (1972) in Chesapeake Bay. Romanov, Ryabinin, Lazareva & Zhidkova (1977) found that copper concentrations in the Aegean Sea varied seasonally, with the winter values higher than those of spring and summer. A pronounced seasonal variation was observed in river discharge to the Baltic Sea from Sweden (Ahl, 1977). Spencer & Brewer (1969) were unable to relate the seasonal variation of copper in the surface waters of the Gulf of Maine to fluctuations in primary productivity either because the rate of uptake was too low or the rate of release was sufficient to mask any measurable effect. Localized fluctuations in copper levels have also been linked with tidal cycles; Dehlinger et al. (1973) and Zirino, Clavell & Seligman (1978a) found copper values to be greatest at low slack and lowest during the flood tide. Some of the variation they observed was associated with salinity changes over the tidal flux. S.E.Jørgensen (1979) has modelled the distribution and effects of heavy metals in aquatic ecosystems and Millero (1975) has presented various models developed to explain ion-water and ion-ion

TABLE IV Copper levels in particulates (from selected references); m., mean; *from a review paper, not from original

Oceanography and marine biology

Location

Level

645

Comment

Reference

Atlantic North Atlantic (surface)

74 mg/kg

Chester & Stoner, 1975*

South Atlantic (surface)

52 mg/kg

”

N.E.Atlantic

112 mg/kg

12 samples—particulates

Chester et al., 1978

(H2O). Pacific China Sea (surface)

107 mg/kg

Chester & Stoner, 1975*

202 mg/kg

”

Indian Ocean Indian Ocean (surface) Coastal Great Britain Conway

2.0–2.8 µg/l

Range of means

Elderfield et al., 1971

Hinkley Power Station

9.5–35 µg/l

Intake

Boyden & Romeril, 1974

3–20 µg/l

Outlet

Europe—Mediterranean Coast Lagoon of Venice, Italy

0.6–1.5 µg/l

Lagoon of Grado, Italy

0.2–1.5 µg/l

Mediterranean, Israel

nd–2.1 µg/l

Barbara et al., 1978

nd=not detected

Roth & Hornung, 1977

(m. 0.4) Europe—Atlantic Coast Bornholm Sea

78–172 µg/g dry wt

Averages from 7 depths ranging from 1 to 72 m

Kremling & Petersen, 1978

Sörfjord, W.Norway

<0.1–2.04 µg/l

8 stations, various depths

Skei et al., 1973

0–0.4 µg/l

3 stations—range over various depths

Weigel, 1977

Local background range

Jan et al., 1978

Baltic Sea Baltic Sea

North America—Pacific Coast Southern California

0.01 µg/kg

Municipal wastewater discharge

0.9–1.4 µg/kg Median values

Interlinking of physical

1120

Southern California Bight

1–410 ng/kg

Franks & Bruland, 1978

Scripps, California

0.33–3.00 µg/l

Ford et al., 1976

(m. 1.08) San Diego, Encina Power Plant

0.33–2.00 µg/kg

Intake

”

Thermal effluent

”

In paniculate layer at 125m

Emerson et al., 1979

(m. 0.91) 0.33–3.24 µg/kg (m. 0.93) Saanich Inlet, B.C.

0.16% by wt

Columbia River Estuary, Washington

130–698 ng/kg

Cutshall et al., 1973

North America—Atlantic Coast Narragansett Bay

0.11–1.7 µg/l Range of means from surface (5 Piotrowicz et al., samples) 1972

Narragansett Bay

≤0.05–0.55 µg/l

Subsurface (5 samples)

0.44 µg/l

Low temperature ashing

0.44 µg/l

Wet digestion

Wallace & Duce, 1975

North America—Coastal Gulf of Mexico West Coast Puerto Rico

0.009 mg/l

Culebrinas

0.007–0.08 mg/l

Añasco

0.006 mg/l

Guanajibo

0.1–8.9 µg/l

Surface

0.2–2.6 µg/l

Bottom

Lowman et al., 1966

North Sea North Sea

Dutton et al., 1973

interactions of the major components of sea water and how sea water affects the state of metal ions. COPPER LEVELS IN MARINE SEDIMENTS

Oceanography and marine biology

647

The ultimate site of copper deposition is in the sediments (Goldberg, 1957). As a result, the concentrations of copper these are generally higher than those found in overlying water (e.g. Goldberg, 1957; Enomoto, Matsui & Uchida, 1972; Presley, Kolodny, Nissenbaum & Kaplan, 1972). The residence time most widely accepted is 5.0×104 years (Goldberg, 1963 in Horne, 1969). (Residence time is the length of time required for all of the element to be removed and replaced by material of continental origin.) Bernhard (1978) has prepared a literature review of heavy metal levels in the Mediterranean while Förstner et al. (1978) provide a review of sources of metals in sediments. Burton (1978b) gives a general review of the modes of association of trace metals with certain components of the sedimentary cycle. There is a certain amount of evidence that metals are refluxed into the water column from the sediments (e.g. Fig. 1, p. 509) although no work deals specifically with copper. Duke, Willis & Wolfe (1967, 1968), using 65Zn, calculated the exchange rate for zinc to be 17±4 µg Zn·h−1·m−2. A general order of displacement of metals from a sediment is Cu>Co>Zn>Mn (e.g. Johnson et al., 1967). As copper displaces several metals and is itself theoretically not displaced, it could be predicted that the refluxing rate of copper is likely to be much less than that which has been demonstrated for zinc. Heggie (1978) conducted a physical-chemical study of reservoirs, fluxes, and pathways in an Alaskan fjord and found that copper was removed from the water column and transported to the sediments by particulate matter. The net annual removal in this fashion was estimated to be between 9.6 and 14.2 µg Cu·cm−2. Copper was re-mobilized from the solid phase(s) in surface sediments and subsequently returned to the overlying water. The net annual transport across the sediment-sea water interface was estimated to be 1.9 µg Cu·cm−2 which means that 13 to 20% of the copper removed from the water column to the sediments was returned to the water column. Duinker et al. (1974) suggested that any mobilization of copper from bottom sediments leads to increased levels in interstitial waters. Pore and sea water in the radiolarian ooze area of the central Pacific was found to have a higher concentration of copper when compared with the literature value for copper in the sea (Schnier et al., 1978). Heggie & Burrell (1978), working in Resurrection Bay, Alaska, found interstitial copper levels ranging from 1.02–9.98 µg·1−1 with maximum concentrations always being in the top 7 cm of sediment. Conversely, Duchart et al. (1973) found that copper was enriched in the interstitial water in the lower portions of 100-cm cores in a reducing environment with high levels of organic material. Presley et al. (1972) showed that the concentration of copper in interstitial waters of an anoxic sediment was 0.5–39 µg·1−1 and was related to the organic content of the sediment. These and other studies suggest that organic matter in the interstitial water may increase the amount of copper (e.g. Bader et al., 1960; Murray, 1973; Schmidt, 1978a, b). Much the same conclusions can be drawn for the sediment particles

TABLE V Selected copper levels in interstitial (pore) water: units are µg/l

Interlinking of physical

Location

Level

1122

Comment

Reference

North America-Pacific Coast Resurrection Bay, Alaska 1.02– 9.98

Heggie & Burrell, 1978

Rupert and Holberg Inlets, 0.15– B.C. 14.4

Area of copper mine tailings

Thompson, 1975

Howe Sound, B.C.

0.79– 25.3

Area of copper mine tailings

Thompson, 1975; Thompson & Paton, 1978

Saanich Inlet, B.C.

0.5–39.0 Organic rich, stagnant fjord

Presley et al., 1972

San Francisco Bay

0–300

Serne, 1977

San Pedro Bay

0.4

Sandy silt

0.9–1.3

Silty sand

0.4

Silty clay

Gupta & Chen, 1975

Coastal Great Britain Conway Estuary

210

Tees Estuary

365

Elderfield et al., 1975

Europe-Atlantic Coast River Rhine Estuary

4–190

Duinker & Nolting, 1978

themselves. It has been suggested that one of the major influences on copper retention by sediments is the presence of organic material (e.g. Anderson, 1975; Hallberg, 1974; Jaffe & Walters, 1977; Paul & Mieschner, 1976; Armstrong et al., 1976; National Marine Fisheries Service, 1976b). Some of this organic material binds to the surface of particulate material, forming site(s) of metal complexation. Reducing conditions arising from microbial activity in sediments with higher organic content may result in dissolution with the release of trace metals. Okutani & Okaichi (1971) show that the concentration of copper in sediments is between 3.9 and 45.2 mg·kg dry wt−1 with the concentration being greatest where the concentration of sulphate-reducing bacteria is greatest. Laube et al. (1979) have proposed accumulation by algae as an important mechanism for mobilizing sediment-bound metal ions. Salomons & Mook (1977) show a decrease in copper values in bottom sediments of the Rivers Rhine and Ems in a seaward direction, suggesting flocculation, possibly with humic or fulvic compounds in the salt-water-freshwater mixing zone. Pheiffer (1971), Hershelman et al. (1977), Hiraizumi, Manabe & Nishimura (1978), and Ramondetta & Harris (1978) discuss the initial distribution of copper in sediments relative to their source and factors affecting the movement(s) of the metalcontaining sediments on initial deposition. The copper content of sediments has been related to sediment size and geochemical make-up (e.g. De Groot, 1970; Naidu & Hood, 1971; Hallberg, 1974; Armstrong et al.,

Oceanography and marine biology

649

1976; Paul & Mieschner, 1976; Jaffe & Walters, 1977; Turekian, 1977; Loring, 1978c; Matsui et al., 1978; McDuffie, El-Barbary, Hollod & Tiberio, 1978, Afran, Khalily & Nickless, 1979; Lyons & Gaudette, 1979b; Simpson, 1979) with the finer sediment fractions being most closely correlated with higher metal values regardless of chemical variables except organic content. Goldberg et al. (1977) as well as others (e.g. Erlenkau, Suess & Willkon, 1974; Gardner, 1976; Chen, 1977;

Amiel & Navrot, 1978; Sommer & Pyzik, 1974; Thompson & McComas, 1974) show a tremendous decrease in copper levels with depth in core samples. The decrease is attributed in many cases to progressive industrial and urban development and subsequent contamination of the more recent sediments (e.g. Dominik et al., 1978; Skei & Paus, 1979). Taylor (1979), in a study of the effects of discharges from three industrialized estuaries on the distribution of heavy metals in the coastal sediments of the North Sea notes, however, that the geology of the area can be a more important factor than the industrial input in deciding the metal content of marine sediments. Other studies point out post-depositional re-distribution under natural conditions resulting from pH-Eh gradients (Bonatti et al., 1971; Hallberg, 1974). Schmidt & Gibson (1978) suggest that sediment organocopper compounds, in a form solubilized by dilute acetic acid (i.e., inorganically precipitated or weakly adsorbed), is the most likely source of copper from amended sediments. COPPER LEVELS IN MARINE ORGANISMS A major review of references concerning the elemental composition of marine organisms is given by Vinogradov (1953). Other major reviews have been prepared by Lunde (1970), Hueck-van der Plas (1972), Won (1973), Bouquiaux (1974), Stenner & Nickless (1975), Gerlach (1976), Leland et al.

TABLE VI Copper levels in marine sediments: units are mg/kg; *a review paper containing levels used in determining the range; this table is for ordering selected references rather than providing detailed levels of copper

Location

Range

References

Atlantic Ocean

48– 163

Chester & Stoner, 1975*; Spencer et al., 1978

Coastal Great Britain

1– 2215

Elderfield et al., 1971; Halcrow et al., 1973; Boyden & Romeril, 1974; Taylor, 1974; Aston & Thornton, 1975; L.H.Jones et al., 1976; Taylor, 1976*; Jaffe & Walters, 1977*; Phillips, 1977*;

Interlinking of physical

1124

Roth & Hornung, 1977*; Bryan & Uysal, 1978; Foster et al., 1978; Hetherington & Harvey, 1978; Jones & Jordan, 1979 Coastal European 2– Piper, 1971; De Groot, 1973; Förstner & Reineck, 1974; Stenner Atlantic including the 12000 & Nickless, 1974a; Müller & Förstner, 1975; Armstrong et al., North and Baltic Seas 1976*; Loring, 1976*; Neelakantan & Kusuma, 1976; Djafari, 1977; Jaffe & Walters, 1977*; Lande, 1977; Patchineelam & Förstner, 1977; Phillips, 1977*; Salomons & Mook, 1977 Mediterranean and Black Seas

Location

0.1– 52.0

Lev-er et al., 1976; Loring, 1976*: Roth & Hornung, 1977*; Griggs & Johnson, 1978; Collinson & Rees, 1979

Range

References

Coastal North <1·94– Bopp & Biggs, 1972; Sommer & Pyzik, 1974; Villa & Johnson, 1974; America 8054 Wood & Cintron, 1974; Chester & Stoner, 1975; Helz et al., 1975; Windom, 1975; Armstrong et al., 1976*; Fink et al., 1976; Loring, 1976; National Marine Fisheries Service, 1976b; Willey, 1976; Wolfe et al., 1976; Greig & McGrath, 1977; Greig et al., 1977*; Jaffe & Walters, 1977*; Roth & Hornung, 1977*; Summerhayes et al., 1977; Goldberg et al., 1978a; Loring, 1978a; Olsen et al., 1978; Pilotte et al., 1978; Timoney et al., 1978 Gulf of Mexico

1·6– 255

Reynolds & Thompson, 1974; Armstrong et al., 1976*; Trefry et al., 1976*; Phillips, 1977*; Roth & Hornung, 1977*; Khalid et al., 1978

Coastal Australia and New Zealand

2·2– 102

Knauer, 1977; Phillips, 1977*

Coastal Northwest Pacific

0·03– 289·1

Inoue et al., 1974b; Kawakami & Nishimura, 1976; Wong & Li, 1977; Yamamoto et al., 1977a; Kurata & Yoshida, 1978; M.H.Wong et al., 1978

Beaufort Sea (Mackenzie River delta)

23–42

Loring, 1976*; Wagemann et al., 1977

Central and Eastern Pacific

18– 5300

Burnett, 1975; Oldnall, 1975; Turner, 1975; Schmidt, 1978b*

Coastal North 2·6– American 1394 Pacific

Galloway, 1972a; Pal, 1974; Thompson & McComas, 1974; Anderlini et al., 1975; Serne & Mercer, 1975; Gupta & Chen, 1975; Armstrong et al., 1976*; Bascom et al., 1976; Bradford, 1976; Ecker & Hendricks, 1976; Lu & Chen, 1977; Phillips, 1977*; Serne, 1977; McGreer, 1979b

(1976), Reish et al. (1977), Schmidt (1977, 1978a, b), Bernhard (1978) Hall, Zook & Meaburn (1978), Leland et al. (1978), Reish et al. (1978), and Eisler et al. (1979). Several problems exist when attempting to compare published metal levels in marine organisms. First, and perhaps most important, are the numerous methods of analysis and

Oceanography and marine biology

651

expression of results (method of extraction and measurement; wet or dry weight; tissue only or whole organism including skins, shells, or exoskeletons; etc.). These, combined with seasonal and geographical variation in metal levels found in the habitat, may yield a wide range of values for any one species. Secondly, are the problems of species identification and variation of metal levels during the life history. We have presented data from selected references; numerous values have been omitted for reasons of clarity, brevity, repetition or oversight. In general, proportions of the minor and trace elements including copper, vary much less than the major elements Na, K, Sr, and SiO2 (Moore & Boström, 1978); although some organisms obviously have the ability to concentrate copper while others do not. Individual cells of the oyster Ostrea edulis have been found to contain as much as 13 mg copper·g−1 with ambient water levels of 0.065 µg Cu·ml−1 (George et al., 1978). Ruddell & Rains (1975) discuss diapedesis in oysters, the shedding of large numbers of zinc and copper rich blood cells called “basophils” to regulate intraorganism metal levels. Seaweeds, on the other hand, have accumulation capabilities for copper that tend to reflect the environmental concentration of the metal. The nature of uptake and accumulation processes have been discussed on pages 507–522. The grouping of organisms in this section has been made with respect to the major life style of the organisms (planktonic, benthic, nektonic). PLANKTONIC ORGANISMS: (Tables VII and VIII) The uptake of soluble copper by planktonic organisms is a primary pathway for entry of metal into the food web; ingestion of abiotic particles, such as clay minerals, may also lead to copper accumulation by marine organisms. Davies (1978) provides an excellent review of the literature and theory of heavy metals and marine plankton. Chester & Stoner (1975) give a range of copper in marine plankton of 12–56 µg·g dry wt−1 while Oppenheimer (1968) gives a concentration of 1100 µg·g ash wt−1 for mixed plankton. Concentrations in the plankton of Long Island Sound were shown to range from <2.0– 39.3 µg·g dry wt−1 (Greig et al., 1977) while Fowler (1977) found a mixture of plankton (principally copepods, phytoplankton, chaetognaths, and detritus) to have a level of 39 µg Cu·g dry wt−1. Due to the diverse nature of the life styles and metal accumulation tendencies and mechanisms which one encounters in a plankton population, the usefulness of copper values for mixtures of organisms is questionable. Measurements of copper levels encountered in individual or defined groups of plankton are far more useful, particularly for considerations of bioaccumulation and food web analysis. Marine Bacteria Copper values presented in the literature for bacteria are somewhat artificial in that the populations have, in many cases, been grown in culture media. Oppenheimer (1968) cultured a variety of bacteria in a defined medium (0·1 mg copper·1−1) and determined the final concentrations of a variety of trace metals. He found a mixed anaerobic bacteria culture to contain 100 µg Cu·g−1, contaminated sulphur reducer 100 µg·g−1, and a pure sulphate reducer 90 µg·g dry wt−1. Jones, Royle & Murray (1978) found an average of 25 µg·g−1 for dried bacterial cells grown in synthetic medium.

Interlinking of physical

1126

Marine Phytoplankton Levels of copper in phytoplankton are quite variable. While the need for copper has been demonstrated, the amount found in the organism is largely a function of the concentration in the sea water in which it exists. Each species also varies in its ability to take up and concentrate metals, some appear to do this passively while others actively regulate their metal concentration. Problems associated with obtaining large enough samples hinder the description of copper levels in the field. Some general levels for algae/phytoplankton have been included in Tables VII and IX as it was not possible to discern the species composition of the samples from the references.

TABLE VII Levels of copper in selected marine phytoplankton: *from a review paper, not from original

Species

Level

Comments

Reference

Plankton

2.4–226.3 µg/g (mean 24.0)

Off N.W.Africa

Brugmann, 1978

Mixed phytoplankton

36 mg/kg dry wt

Irish Sea

Riley & Roth, 1971

Phytoplankton

25 mg/kg dry wt

Netherlands

Reish et al., 1977*

24 mg/kg dry wt

Off N.W.Africa

Brugmann, 1978

Coscinodiscus jonesianus

1.3–1.9×10−3 mg/kg dry wt

Sea of Azov

Rozhanskaya, 1967

C.curvisetus

0.5–1×10−1% ash wt

Black Sea

Vinogradova, 1965

Leptocylindricus danicus

9×10−2 % ash wt

Sea of Azov

”

Bacillariophyceae

Rhizoselenia calcaravis 1×10−2 % ash wt

Black Sea

”

5×10−3% ash wt

Caspian Sea

”

1–2×10−2% ash wt

Scotia Sea, Antarctica

”

35 µg/g ash

Mean; Puerto Rico

Ting & de Vega, 1969

Chlorophyceae Udotea flabellum

Marine Zooplankton Vinogradova (1965) states “in common species of zooplankton, copper is concentrated to

Oceanography and marine biology

653

a higher degree (from 3×10−2 to 3×10−1%) than in common forms of phytoplankton (from 1×10−3 to 8×10−2% [ash weight])” suggesting some importance of the relative position of the organism in the food chain. Other factors such as size and ambient copper levels influence concentrations in some benthic and planktonic organisms. Greig et al. (1977) for example, found that metal concentrations varied with geographical location and species in the New York Bight and Long Island Sound areas while Melluso et al. (1978) found a connection between copper levels in the sessile tunicate Ciona intestinalis in the Gulf of Naples and the sediment to which the metal was adsorbed. Leland et al. (1978) give values of 3–61 µg·g dry wt−1 for a mixed zooplankton population of marine and estuarine origin while Greig et al. (1977) found levels of <1.6–54.4 µg·g dry wt−1 for zooplankton in the New York Bight area. The species composition of the population is very important as, in Greig et al. (1977), the minimum value (<1.6 µg·g−1) proved to be for a sample consisting of ctenophores and the maximum value (54.4 µg·g−1) for a sample consisting of 95% rock crab

TABLE VIII Levels of copper in selected marine zooplankton: units unless stated otherwise are mg/kg dry wt

Species Zooplankton

Level

Comments

Reference

<1.0–8.1 µg/kg wet wt

Mid-Atlantic Bight, 3 stations

Palmer & Lear, 1973

<1.6–54.4 mg/kg (most 2.9–22 mg/kg)

New York Bight

Greig et al., 1977a

<2.0–39.3 mg/kg (most 2–14 mg/kg)

Long Island Sound

Estuarine tidal front in Delaware Bay

Sick et al., 1978

7.4±0.4

south of front

17.9±0.9

in frontal zone

2.1±0.8

north of front

2.4–226.3

Off N.W.Africa

Brugmann, 1978

68

Gullmarfjord

Noddack & Noddack, 1939

Coelenterata Cyanea capillata

Physalia sp.

8.2–13 mg/kg ash Continental Shelf, N.E.United States

Nicholls et al., 1959

2.2–3.7

Rose & Bodansky,

Interlinking of physical

1128

1920 Ctenophora Beröe cucumis

700 mg/kg ash

Continental Shelf, N.E.United States

Nicholls et al., 1959

30 mg/kg ash

Continental Shelf, N.E.United States

„

33.5–107

Greece

Reish et al., 1978*

3.4–3.7

Canadian Arctic

Acartia clausi

34.0–107

Elefsis Bay, Greece

Zafiropoulos & Grimanis, 1977

Anomalocera patersoni

41.9

Mediterranean Sea

Polikarpov et al., 1979

Pseudodiaptomus coronatus and Acartia tonsa (mixed)

19

Georgia and S.Carolina

Stickney et al., 1975

Neomysis americana

25±11

Georgia and S.Carolina

”

Euphausiids

48

N.W.Mediterranean

Fowler, 1977

Euphausia krohnii

600 mg/kg ash

Continental Shelf, N.B.United States

Nicholls et al., 1959

Meganyctiphanes norvegica

19–44 mg/kg

Firth of Clyde, U.K.

Halcrow et al., 1973

Amphipods

35–94

Norway

Stenner & Nickless, 1974a

26

Canadian Arctic

Reish et al., 1978*

Gammarus sp.

30

Georgia and S.Carolina

Stickney et al., 1975

Chaetognatha

5.6–6.3

Canadian Arctic

Reish et al., 1978*

Sagitta elegans

1100 mg/kg ash

Continental Shelf, N.E.United States

Nicholls et al., 1959

Sagitta sp.

5.0

Clyde Estuary, Scotland

McIntyre, 1976*

500 mg/kg ash

Continental Shelf, N.E.United States

Nicholls et al., 1959

Mollusca (Pteropoda and Heteropoda) Limacina retroversa Crustacea Copepods

Larvacea Salpa fusiformis

larvae. As copper levels vary with species many of the existing values should be treated with caution. Polikarpov et al. (1979) found that the body burden of copper in the neustonic copepod Anomalocera patersoni collected from the Mediterranean showed no

Oceanography and marine biology

655

variation with respect to sex. Very few publications provide copper concentrations in individual species of planktonic polychaetes and ostracods. In general, there is relatively little information on copper levels in marine zooplankton and much of it is for major groups or just for “zooplankton”. BENTHIC ORGANISMS (Tables IX and X) Algae Measurements of copper in seaweed range from 6–300 µg·g−1 (Schmidt, 1978a,b). Algae containing >50 µg Cu·g−1 appear to be either particularly effective accumulators or have been exposed to abnormally high environmental concentrations (or both). Numerous studies suggest that increased metal concentrations in the environment are due to industrial and urban activity and result in increased levels in algae. Examinations of Ulva sp. and Enteromorpha sp. from Iona Island sewage outfall and from Roberts Bank (less polluted) in British Columbia (Canada) coastal waters showed the levels at Iona Island to be about an order of magnitude greater (Vermeer & Peakall, 1979). Agadi, Bhosle & Untawale (1978) discuss the rôle of different seaweeds as indicators of marine pollution. The age and type of tissue analysed play an important part in determining algal metal levels as older tissues generally contain higher metal concentrations. Fucus vesiculosus, sampled in the Tamar Estuary, U.K., by Bryan (1971), contained more copper in old thalli than in newly developed tissue. These findings are also reflected in the work on F.distichus by Bohn (1979) who found copper concentrations to be higher in older tissue, highest in the stipe. Ishii, Suzuki & Kojanayi (1978) found variations in copper concent ration from 1.4–2.2 µg·g dry wt−1 dependent upon which organ of Sargassum horneri they analysed, or 1.6–2.0 µg·g dry wt−1 using different growth stages. Copper values were found to nearly double from the roots to the leaves in various sea grasses (Pulich et al., 1976) while Haug et al. (1974) showed a steady increase in copper from the receptacle through the tips and internodes 1–4 of Ascophyllum. Bryan (1969) found variations related to age within individual plants of the genus Laminaria. Invertebrates Metal levels reported in the literature for invertebrates appear dependent upon ambient conditions, the age and weight of the organism, the specific tissues examined, and the overall physiological capabilities of the organism with respect to metal uptake. Copper levels in invertebrates quoted in this review range up to 15 160 µg·g dry wt−1 found in the liver of a squid (Loligo opalescens) from central California (Martin & Flegal, 1975). Most authors, working with a wide range of invertebrates, accept the idea that the extent of copper uptake by organisms is directly related to the concentration to which they are exposed. The nature of this relationship,

TABLE IX

Interlinking of physical

1130

Levels of copper in marine benthic algae: units unless stated otherwise are mg/kg dry wt; m., mean. *from a review paper, not from original

Species

Level

Comments

References

Chlorophyceae Chlorophyceae

“Green algae”

11 mg/kg ash wt

Puerto Rico

Lowman et al., 1966

1.2–21

Vostok Bay, Sea of Japan

Phillips. 1977*

15–170

New York

Reish et al., 1978*

0.43–3.3 mg/kg S.Africa wet wt

”

2.9–5.5

Israel

”

11–34

England

Reish et al., 1977*

Bryopsis

5.5

Newport River estuary, N.C.

Wolfe et al., 1976

Caulerpa peltata

12.59

Goa coast, India

Agadi et al., 1978

C.racemosa var. laetevirens

6.85

Penang Island, Malaysia

Sivalingam, 1978

C.sertularioides

3.2–12.6

Goa coast, India

Zingde et al., 1976

5.03

Goa coast, India

Agadi et al., 1978

Cladophora fascicularia

7.40

Penang Island, Malaysia

Sivalingam, 1978

Codium elongatum

3.82

Goa coast, India

Agadi et al., 1978

Enteromorpha clathrata

30.23

Goa coast, India

E.flexuosa

26.12

Penang Island, Malaysia

Sivalingam, 1978

E.intestinalis

60.1–61.8 (m. 60.9)

Restronguet Creek, U.K.

Klumpp & Peterson, 1979

Enteromorpha sp.

19–110

Hardanger Fjord, and Skjerstad Fjord, Norway

Stenner & Nickless, 1974a

7.90 mg/kg wet Iona Island, B.C., Canada wt

Vermeer & Peakall, 1979

0.78 mg/kg wet Robert Banks, B.C., Canada wt

Halimeda tuna

6, 18, 22

S.W.Atlantic coast, Spain/Portugal, 3 locations

Stenner & Nickless, 1975

15.0–61.9

Coast of Ras Beirut, Lebanon

Shiber & Shatila, 1979

Oceanography and marine biology

Ulva fasciata

U.lactuca

657

6.70–7.30

Goa coast, India

Agadi et al., 1978

6.8–12.4

Goa coast, India

Zingde et al., 1976

11–34

Poole Harbour, U.K.

Phillips, 1977*

5.5–26

Atlantic coast of Spain and Portugal

”

9–170

Skjerstad Fjord, Norway

Stenner & Nickless, 1974a

3000 mg/kg ash wt

Vinogradov, 1953*

5.5–26

S.W.Atlantic coast, Spain– Portugal, 4 locations

Stenner & Nickless, 1975

U.pertusa

12

Japan

Ishii et al., 1978

U.reticulata

4.49–12.8

Singapore coast

Bok & Keong, 1976

13.85

Goa coast, India

Agadi et al., 1978

Species Ulva sp.

Level

Comments

References

m. 20

Newport River estuary, N.C.

Wolfe et al., 1976

8.28 mg/kg wet wt

Iona Island, B.C., Canada

Vermeer & Peakall, 1979

0.85 mg/kg wet wt

Roberts Bank, B.C., Canada

”

Valonia fastigiata

16.69

Penang Island, Malaysia

Sivalingam, 1978

Valoniopsis pachynema (young)

11.05

Penang Island, Malaysia

”

V.pachynema (matured)

Trace

Penang Island, Malaysia

”

3.78

Penang Island, Malaysia

”

Rhodophyceae, 4 sp.

4.8–12.9

Vostok Bay, Sea of Japan

Phillips, 1977*

“Red algae”

7.6

Israel

Reish et al., 1978

Acanthophora orientalis Trace

Penang Island, Malaysia

Sivalingam, 1978

A.specifera

7.40–80.41

Goa coast, India

Agadi et al., 1978

Agardiella sp.

m. 8.7

Newport River estuary, N.C.

Wolfe et al., 1976

Chondrus crispus

6

Cyanophyceae Pelagothrix clevei Rhodophyceae

8

Vinogradov, 1953* S.W.Atlantic coast, Spain/Portugal, 1 location

Stenner & Nickless, 1975

Interlinking of physical

Corallina officinalis

1132

11

Skjerstadfjord, Norway

Stenner & Nickless, 1974a

7.5–8

S.W.Atlantic coast, Spain/Portugal, 2 locations

Stenner & Nickless, 1975

Delesseria sanguina

9

S.W.Atlantic coast, Spain/Portugal, 1 location

”

Dasya sp.

m. 17

Newport River estuary. N.C.

Wolfe et al., 1976

Gigartina mamillosa

15

Norway

Lunde, 1970

5.6–6.6

Norway

Oy, 1940

G.stellata

5–33

Hardanger and Skjerstad fjords, Norway

Stenner & Nickless, 1974a

Gracilaria corticata

3.52–9.22

Goa coast, India

Agadi et al., 1978

G.verrucosa

15.61

”

”

Gracilaria sp.

m. 3.2

Newport River estuary, N.C.

Wolfe et al., 1976

Gracilaria sp. 1

Trace

Penang Island, Malaysia

Sivalingam, 1978

Gracilaria sp. 2

10.53

Penang Island, Malaysia

”

Gracilaria sp. 3

2.84

Penang Island, Malaysia

”

Gymnogongrus sp.

m. 5.6

Newport River estuary, N.C.

Wolfe et al., 1976

Species

Level

Comments

References

Halosaccion ramentaceum

3.2

Iceland, 31.8‰ salinity

Munda, 1978

Hypnea musciformis

8.6–22.3

Goa coast, India

Zingde et al., 1976

6.55–8.56

Goa coast, India

Agadi et al., 1978

Hypnea sp.

Trace

Penang Island, Malaysia

Sivalingam, 1978

Jania sp.

4.50

Penang Island. Malaysia

”

Laurencia sp. 1

13.43

Penang Island, Malaysia

”

Laurencia sp. 2

4.74

Penang Island, Malaysia

”

Palmaria palmata

3.0

Iceland, 32.0‰ salinity

Munda, 1978

Porphyra purpurea

80

Iceland, 12‰ salinity

”

P.umbilicalis

2.8–23.3

Coast of Britain

Phillips, 1977*

Porphyra spp.

5.9–67

North Sea shoreline, England

Dutton et al., 1973

Porphyra sp.

20

Iceland, 32.9‰ salinity

Munda, 1978

Pterocladia pinnata

11.3–116.9

Coast of Ras Beirut, Lebanon

Shiber & Shatila, 1979

Oceanography and marine biology

Rhodymenia palmata 33.5

659

Norway

Oy, 1940

24

Norway

Lunde, 1970

5.20

Penang Island, Malaysia

Sivalingam, 1978

Phaeophyceae, 8 sp.

0.9–4.3

Vostok Bay, Sea of Japan

Phillips, 1977*

“Brown algae”

12–38

New York

Reish et al., 1978*

<2–231

England

”

3.5–33

England

”

6–123

Norway

”

Alaria esculenta

4.9

Norway

Oy, 1940

Ascophyllum nodosum

4–85

Hardangerfjord and Skjerstadtfjord, Norway

Stenner & Nickless, 1974a

Sarcodia sp. Phaeophyceae

6–18

Menai Straits, U.K.

Phillips, 1977*

46–96

Dulas Bay, U.K.

”

11–125

Sorfjorden, Norway

Melhuus et al., 1978

4–8 (m. 5.5)

Norway, Reine in Lofoten (not industrialized)

Haug et al., 1974

4–240

Trondheimsfjord (industrialized)

”

3–160

Hardangerfjord (industrialized)

”

1.1–1.4

Norway

Oy, 1940

18–63

Norway

Lunde, 1970

3.4

Iceland, 32.7‰ salinity

Munda, 1978

6.3

12.5‰ salinity

„

89.4–228.3 (m. Restronguet Creek, U.K. 96.2)

Species Chorda filum

Level

Coments

Klumpp & Peterson, 1979

References

7–100

Hardangerfjord and Skjerstadtfjord, Stenner & Nickless, Norway 1974a

m. 10–16.7

Sorfjorden, Norway, 4 stations

5.5

S.W.Atlantic coast, Spain/Portugal, Stenner & Nickless, 1 location 1975

Chordaria flagelliformis

5.6

Iceland, 32.7‰ salinity

Munda, 1978

Colpomenia sinuosa

4.50

Penang Island, Malaysia

Sivalingam, 1978

Melhuus et al., 1978

Interlinking of physical

1134

21.06

Goa coast, India

Agadi et al., 1978

22.5–52.5

Coast of Ras Beirut, Lebanon

Shiber & Shatila, 1979

Dictyopteris australis 8.36–9.72

Goa coast, India

Agadi et al., 1978

Dictyosiphon chordaria

8.8

Iceland, 29.7‰ salinity

Munda, 1978

D.foeniculaceus

5.6

Iceland, 32.7‰ salinity

”

Dictyota bartayresii

49.74

Penang Island, Malaysia

Sivalingam, 1978

D.dumosa

12.9

Goa coast, India

Zingde et al., 1976

Dictyota sp.

8.9

Goa coast, India

”

Dictyota spp.

5.84–12.99

Goa coast, India

Agadi et al., 1978

Ectocarpus sp.

m. 9.4

Newport River estuary, N.C.

Wolfe et al., 1976

Egregia menziesii

0.4–4.9

S.California Bight

Callahan & Stokes, 1978

Fucus ceranoides

21–30

Hardangerfjord, Norway

Stenner & Nickless, 1974a

Iceland,

Munda, 1978

2.8

18.0‰ salinity

3.0

4.2‰ salinity

F.distichus

Strathcona Sound NWT pre-mining Bohn, 1979 of Pb-Zn ore, 2.5–4.7

0.1 km downstream of drainage

(m. 3.3) 2.3–3.3 µg/g dry wt

18 km downstream

”

Munda, 1978

(m. 3.0) F.distichus ssp. edentatus

1.8

Iceland, 31.9‰ salinity

F.serratus

5–26

Hardangerfjord and Skjerstadtfjord, Stenner & Nickless, Norway 1974a

5.9–204

La Rosiere, below desalination plant Romeril, 1977

<2.5–8.1

Pont LeFret control

”

m. 12–15

Sorfjorden, Norway; 3 stations

Melhuus et al., 1978

19

Power Station outfall;

Boyden & Romeril, 1974

18

intake; Hinkley, U.K.

Oceanography and marine biology

5.8–17.4

661

Norway

Oy, 1940

30.0

Species F.serratus

Vinogradov, 1953*

Level

Comments

References

8.0–27

Norway

Lunde, 1970

62.4– 68.2

Restronguet Creek, U.K.

Klumpp & Peterson, 1979

Hardangerfjord, Norway

Stenner & Nickless, 1974a

(m. 65) F.spiralis

12

4.1–231 La Rosiere, below desalination plant

Romeril, 1977

<2.0– 3.75

Pont LeFret control

12–43

Norway

Lunde, 1970

71.3– 86.2

Restronguet Creek, U.K.

Klumpp & Peterson, 1979

”

”

Hardangerfjord and Skjerstadtfjord, Norway

Stenner & Nickless, 1974a

(m. 78.6) F.vesiculosus

86.2– 91.3 (m. 89) 4–118

1.6–36.0 Coast of Britain

Phillips, 1977*

9–301

”

4 estuaries, England

2.8–14.3 Bristol Channel, U.K.

”

1.7–6.6

Caernarvon Bay, U.K.

”

49–97

Dulas Bay, U.K.

”

17–150

Sorfjorden, Norway

Melhuus et al., 1978

3.4–8.4

Norway

Oy, 1940

20–107

Tamar estuary, U.K.

Bryan & Uysal, 1978

12–45

Norway

Lunde, 1970

Iceland,

Munda, 1978

3.0

32.1‰ salinity

3.4

32.7‰ salinity

4.0

29.0‰ salinity

Interlinking of physical

Fucus sp.

1136

3.6

12.5‰ salinity

9–31

Atlantic coast of Spain and Portugal

9 & 31

S.W.Atlantic Coast, Spain/Portugal, 2 Stenner & Nickless, 1975 locations

5–10

Netherlands

Hueck, 1975

2.5–24

North Sea shoreline England

Dutton et al., 1973

3.5–33

Cornwall, England

Bryan & Hummerstone, 1977

6–36

Hardangerfjord and Skjerstadtfjord Norway

Stenner & Nickless, 1974a

1.3–4.0

England, 9 locations

Phillips, 1977*

Phillips, 1977*

(m. 7.8)

Laminaria digitata

3.2–10.2 Norway

Oy, 1940

14–34

Norway

Lunde, 1970

L.gloustoni

2.4–4.8

Norway

Oy, 1940

L.hyperborea

6–22

Norway

Lunde, 1970

L.saccharina

3.2–12.4 Norway

Oy, 1940

L.stilker

4.8–5.5

Norway

”

Macrocystis pyrifera

0.9–4.1

S.California Bight

Callahan & Shokes, 1978

Macrocystis sp.

1.46– 1.47

S.California Bight

Science Applications Inc., 1978c

Padina commersonii

3.78– 7.28

Singapore coast

Bok & Keong, 1976

Species

Level

Comments

P.tenuis

5.69

P.tetrastromatica

8.7–20.1 Goa coast, India

Zingde et al., 1976

4.03– 7.86

Goa coast, India

Agadi et al., 1978

8–41

Norway

Lunde, 1970

62.3– 69.8

Restronguet Creek, U.K.

Klumpp & Peterson, 1979

Pelvetia canaliculata

Penang Island, Malaysia

References Sivalingam, 1978

(m. 67.6) Sargassum grevillei

5.15

Penang Island, Malaysia

Sivalingam, 1978

S.horneri

1.4–2.2

Plant organs

Ishii et al., 1978

Oceanography and marine biology

663

1.6–2.0

Growth stages

S.kjellmanianum

10

Japan

”

S.ringgoldianum

9.9

Japan

”

S.sagamianum

6.9

Japan

”

S.tenerimum

10.3– 27.5

Goa coast, India

Zingde et al., 1976

4.53– 15.46

Goa coast, India

Agadi et al., 1978

S.thunbergii

5.9

Japan

Ishii et al., 1978

Sargassum sp.

3.28– 7.29

Singapore coast

Bok & Keong, 1976

Spathoglossum asperum 5.54

Goa coast, India

Agadi et al., 1978

Stoechospermum marginatum

Goa coast, India

„

33.60

Others—plants, grasses, reeds Amphibolis antarctica

1.3–7.3

Western Port, Australia

Harris et al., 1979

Carex lyngbyei

94.3

Oregon, initial natural marsh

Wolf et al., 1978

70.4

Control plants at 8 wk

4400.0

Experimental plants at 8 wk (exposed to dredge material)

Deschampsis cespitoma 9.1

Distichlis spicata

Oregon, initial natural marsh

”

17.3

Control plants at 8 wk

”

4100.0

Experimental plants at 8 wk (exposed to dredge material)

”

6.85

Georgia, initial natural marsh

”

7.10

Control plants at 8 wk

3840.0

Experimental plants at 8 wk (exposed to dredge material)

76.5

Oregon, initial natural marsh

29.8

Control plants at 8 wk

3980.0

Experimental plants at 8 wk (exposed to dredge material)

Halodule

Redfish Bay, Texas, leaves

Pulich et al., 1976

6.0 4.5 Halophila

roots Redfish Bay, Texas, leaves

”

Interlinking of physical

1138

<2.0

Heterozostera tasmanica

Species Salicornia virginica

<2.0

roots

4.1–6.2

Western Port, Australia

Level

Harris et al., 1979

Comments

7.49

Georgia, initial natural marsh

8.20

Control plants at 8 wk

3540.0

Experimental plants at 8 wk (exposed to dredge material)

86.9

Oregon, initial natural marsh

39.7

Control plants at 8 wk

4480.0

Experimental plants at 8 wk (exposed to dredge material)

References Wolf et al., 1978

Scirpus paludusus

5.23 mg/kg Iona Island, B.C., Canada; seeds Vermeer & wet wt Peakall, 1979

S.validus

5.99 mg/kg Iona Islands, B.C., Canada: wet wt seeds

”

Spartina alterniflora

0.5–7.8

Saltmarsh, S.E.U.S.A.

Windom, 1975

S.alterniflora, S.patens, Distichus spicata (leaves)

2–6

N.Carolina and Louisiana coast

Anonymous, 1977

S.alterniflora

4.32

Georgia, initial natural marsh

Wolf et al., 1978

21.10

Control plants at 8 wk

2760.0

Experimental plants at 8 wk (exposed to dredge material)

5.14

Georgia, initial natural marsh

7.50

Control plant at 8 wk

4510.0

Experimental plants at 8 wk (exposed to dredge material

Spartina sp.

3.9

S.E. Coast, U.S.A.

Gardner et al., 1976*

Syringodium filiforme

25–32

Alaska

Burrell & Schubel, 1977*

Thalassia testudinum

7.2–72

”

”

Thalassia

25 mg/kg ash wt

Biscayne Bay, Florida; blades

Oppenheimer, 1968

5.0

Redfish Bay, Texas, leaves

Pulich et al., 1976

S.patens

”

Oceanography and marine biology

3.7 Triglochin maritima

665

roots

2.22 mg/kg Iona Island, B.C., Canada wet wt

Vermeer & Peakall, 1979

0.60 mg/kg Roberts Bank, B.C., Canada wet wt Zostera marina

m. 7.9

Newport River estuary, N.C.

Wolfe et al., 1976

7.5

Alaska

Burrell & Schubel, 1977*

Z.muelleri

2.4–10.6

Western Port, Australia

Harris et al., 1979

Zostera sp.

27

Hardangerfjord Norway

Stenner & Nickless, 1974a

9–1350

S.W.Atlantic coast, Spain— Portugal, 3 locations

Stenner & Nickless, 1975

however, varies widely (see pp. 509–522). The polychaete worm, Nereis diversicolor seems to absorb copper directly from the interstitial water of the sediment, the extent of its uptake being directly related to the concentration of copper (Bryan, 1971; Bryan & Hummerstone, 1971), while the polychaete Cirriformia spirabranchia accumulates copper to a particular concentration without regard for ambient levels (Milanovich, Spies, Gunam & Sykes, 1976). Hill & Helz (1973) show that copper uptake by oysters is related to the environmental concentration. Chang (1965) found that in polluted waters copper in oysters may be as high as 1180 µg·g ash wt−1 compared with levels in normal oysters of 5.6 µg·g ash wt−1. The highest concentrations of copper in mussels analysed by Fowler & Oregioni (1976) were found in samples taken from the heavily populated and industrialized port areas of the northwestern Mediterranean, while Alexander & Young (1976) showed levels of copper in digestive glands of Mytilus californianus in the vicinity of waste discharge to be 2–3 times those observed in specimens from unpolluted areas. The effect of environmental levels is also shown by the work of Mathews, Boyne, Davis & Simmons (1979) on oysters from South Carolina; Stephenson (1978) using four mollusc species of different trophic levels in New Zealand; Young & Alexander (1977) and Young et al. (1978) with Mytilus edulis in harbour areas; Hung, Li & Wu (1975) on crabs from Kaosiung Harbour, Taiwan; Eisler et al. (1977b) with the clam Pitar morrhuana in the vicinity of the outfall from a Rhode Island electroplating plant; Khristoforova, Bogdanova & Obukhov (1979) in bivalve molluscs from the islands of the tropical zone of the Pacific; and Harrison & Rice (1979) using oysters. Reynolds (1979) did, however, point out that Arctica islandica did not show this when sampled in the U.S. mid-Atlantic region. He suggests that this variation could be attributed to interspecific differences in feeding habit, motility, or metabolic utilization of the metal or perhaps differential speciation of the metal itself. An interesting correlation of copper content of Crassostrea virginica and environmental levels in southern Chesapeake Bay was shown by Bender & Huggett (1977) who recorded environmental and tissue levels before and after the tropical storm Agnes. They found a definite shift in the concentration of copper

Interlinking of physical

1140

in both the oysters and the environment. Flegal (1978) used simple linear correlation coefficients and multiple analysis of variance on data collected from Acmaea scabra in California, finding a general independence of the elemental metal concentrations from each other and from other biological and geographical variables in the total sample. McGreer (1979b) showed a wide range of tissue to sediment ratios (1.95–7.98) for Macoma balthica collected in the Fraser River Estuary (Canada) and Luoma & Bryan (1978) found no apparent relationship between sediment levels and levels found in M.balthica and Scrobicularia plana in a variety of estuaries. Transport membranes in respiratory and digestive organs are identified by Harrison (in Templeton, 1978) and Young (1978) as the primary sites of copper absorption and accumulation in all organisms. Data presented in Table X illustrate that significantly greater levels of copper are found in these specific tissues or organs in invertebrates. J.S.Young et al. (1979b) show a threshold concentration for copper in Eudistylia vancouveri of 3–6 µg·1−1 during winter conditions with the branchial region probably being the major absorptive site. MacKay et al. (1975b) show that copper concentrations

TABLE X Levels of copper in selected marine invertebrates: units unless stated otherwise are mg/kg dry wt; m., mean; SD, standard deviation; *from a review paper, not from original

Species

Level

Comments

References

Porifera Choanites ficus

5

Vinogradov, 1953*

Unid. sponge

17.1–26.5 mg/kg S.California Bight, whole

Callahan & Shokes, 1978

(m. 21.8) Coelenterata Actinosa metridium

1

Alcyonium digitatum

9.7

Anthea cereus

23.5 mg/kg wet wt

Anthopleura elegantissima

2.8–6.4 mg/kg

Vinogradov, 1953* Irish Sea

Riley & Segar, 1970 Vinogradov, 1953*

S.California Bight, whole

Callahan & Shokes, 1978

Japan

Matsumoto et al., 1965

(m. 4.8) Aurelia sp.

0.03 mg/g ash

Briarum sp.

4.5

Vinogradov, 1953*

Cassiopea xamachana

10.5

”

Eunecia crassa

7

”

Oceanography and marine biology

667

E.rissoa

9

”

Gorgonia acerosa

6

„

Gorgonium flabellum

13

„

Pseudoplexaura crassa 11

”

Physallia sp.

2.2 mg/kg wet wt

”

Plexaura homomolla

30

,,

P.flexulosa

12

”

Poccillopora sp.

2

”

Rhizostoma pulmo

m. 0.26, SD 0.09 From Belgian inshore fishery

De Clerck et al., 1979

Tealia felina

57

Riley & Segar, 1970

Xiphigorgonia anaps

7

Vinogradov, 1953*

33.6

„

Irish Sea

Bryozoa Plumatella fungosa Annelida Eudistylia vancouveri

3.64–7.88

Sequim Bay

J.S.Young et al., 1979b

Polychaetes

6.8–69.8 µg/g wet wt

California

Reish et al, 1977*

1.6–78

England

Reish et al., 1978*

Eisenia bicyclis

3.2–11.9

Japan

Ishii et al., 1978

Eisenia sp.

0.74–0.94

S. California Bight

Science Applications Inc., 1978c

Nephthys hombergi

45

S.W. England estuaries

Bryan & Hummerstone, 1971b

Nephthys sp.

8.9

St. Croix River, Maine

Fink et al., 1976

Nereis diversicolor

39.7–85.6

Hessle, Germany

Jones et al., 1976

40.0–109

Skeffling, Germany

”

2.1–5.1 mg/kg wet wt

Medway estuary, U.K.

Wharfe & Van Den Broek, 1977

22–79

Looe estuary, U.K.

Bryan & Hummerstone, 1977

30.8–84.4

St Croix River, Maine

Fink et al., 1976

N. virens

12.1–15.6

,,

,,

Nereis sp.

42–27 mg/kg wet Chocolate-Jones Bay wt

Guthrie et al., 1979

Interlinking of physical

Species Sand worms (Abarenicola pacifica and Hemipodus borealis

Tubifex sp.

Level

1142

Comments

References

20

Roberts & Sturgeon

Banks, B.C.Canada Bawden et al., 1973

19–20

Cowichan Bay, Canada

„

7–9

Texas. 0‰ salinity

Neff et al., 1978

Mollusca Molluscan tissues

0.1–40.0 mg/kg Coastal U,S,A,, 198 sites

Hall et al., 1978

Pelecypods

0–2498

Reish et al., 1977*. 1978*

Ampelisca milleri

av. 60 mg/kg

San Francisco, U.S.A.

Anderlini et al., 1975

Anadana subcrenata

2.69–5.34

Ariake Sea

Enomoto & Toyodome, 1978

Anodonta sp.

3

Irish Sea

Segar et al., 1971

Arctica islandica

m., 4.44–13.23 Muscle, mid-Atlantic U.S.A. Reynolds, 1979 m., 0.58–2.62

Mid-Atlantic coast, U.S.A.

Wenzloff et al., 1979

Barnea dilatata japonica

19.58–98.25

Ariake Sea

Enomoto & Toyodome, 1978

Brachydontes variabilis

26.3–185.6

Ras Beirut, Lebanon

Shiber & Shatila, 1978

Cardium edule

11

Irish Sea

Segar et al., 1971

6–26

S.W. Atlantic

Stenner & Nickless, 1975

19.89

Spain

Establier, 1978

0.8 mg/kg wet wt

Shoeburyness, Essex, U.K. (natural, relatively uncontaminated)

Howard & Nickless, 1978

5.2–27.2

Looe Estuary, England

Bryan & Hummerstone, 1977

4–9

Poole Harbour, England

Boyden, 1975

2.3 mg/kg wet wt

English Channel

Bryan, 1971

Cerastoderma edule

(m. 9.8)

Chlamys opercularis

240 mg/kg wet Renal organs wt 4.0 mg/kg wet wt

Plymouth, Devon, England (natural, relatively

Howard & Nickless, 1978

Oceanography and marine biology

669

uncontaminated Crassostrea angulata

15.48–327.58 mg/ 100 g dry wt

Spain

Establier, 1969a, b

147.21– 2083.28

,,

Establier, 1978

3–48 mg/kg wet wt

Sydney, Australia

MacKay et al., 1975b

C.gigas and Ostrea angasi m., 19–124 mg/kg wet wt

Tasmanian waters (19 locations)

Thrower & Eustace, 1973

C.gigas

1760–6480

Hinkley, England

Boyden & Romeril, 1974

200–340

Menai Strait, U.K.

”

110–400

Sturgeon and Roberts Banks, Bawden et al., 1973 B.C., Canada

335–1527

Tamar River, Tasmania

Ayling, 1974

m. 200

Poole Harbour, U.K.

Boyden, 1975

C.commercialis

7.8–37.5 mg/kg West coast, U.S.A. wet wt

Pringle et al., 1968

Whole

Martin, 1979

86±30 µg/g

Tamales Bay, CA

860±287 µg/g

Redwood City, CA Gills

178±58 µg/g

Tamales Bay, CA

1910±430 µg/g Redwood City, CA

Species C.glomerata

Level 4.4–380 µg/g wet wt (m. 40.0)

C.madrasensis

C.virginica

Comments

References

New Zealand

Nielsen & Nathan, 1975

Cochin region, India,

Sankaranarayanan et al., 1978

10–140 µg/g

Spat

40–160 µg/g

Juveniles

70–205 µg/g

Adults

6–10, 161 mg/kg wet wt

Vinogradov, 1953*

m; 48–261

Georgia coast (10 stations)

Windom & Smith, 1972

78–230

Coastal U.S.A.

Pringle et al., 1968

Interlinking of physical

1144

7.0–517 mg/kg wet East coast, U.S.A. wt

”

339–630 mg/kg freeze dry wt

Thames River estuary, Conn.

National Marine Fisheries Service, 1976b

9.39–40.66 mg/kg wet wt

Galveston, Texas

Guthrie et al., 1979

Whole

Martin, 1979

273±117 1510±702

Tamales Bay, CA Redwood City, CA Gills

Ensis ensis

Glycymeris glycymeris

295+131

Tamales Bay, CA

1810+372

Redwood City, CA

m. 56.7

South Carolina coast

Mathews et al., 1979

243–4304

Long Island Sound, U.S.A.

Dehlinger et al., 1974

9.4–192.2 mg/kg wet wt

Tasmania

Ratkowsky et al., 1974

109 mg/kg ash wt

Saronikos Gulf, Greece Papadopoulou & Kanias, 1976

99 mg/kg ash wt

Muscle

114 mg/kg ash wt

Mantle and gills

123 mg/kg ash wt

Stomach and intestines

157 mg/kg ash wt

Foot

5.7

Irish Sea

356 mg/kg ash wt

Saronikos Gulf, Greece Papadopoulou & Kanias, 1976

334 mg/kg ash wt

Muscle

275 mg/kg ash wt

Mantle and gills

115 mg/kg ash wt

Stomach and intestines

Gryphea angulata 5.15 mg/kg wet wt Hinnites giganteus 4.3–43.7 mg/kg

Segar et al., 1971

Vinogradov, 1953* S.California Bight, Digestive gland

Callahan & Shokes. 1978

Palos Verdes, CA

Young & Jan, 1976

Digestive gland

Jan et al., 1977

(m. 24) H.multirugosus

Oceanography and marine biology

170 mg/kg wet wt

Wastewater outfall

48 mg/kg wet wt

Control

671

Gonad 3.3 mg/kg wet wt

Wastewater outfall

1.9 mg/kg wet wt

Control Adductor muscle,

0.29 mg/kg wet wt

Wastewater outfall

0.11 mg/kg wet wt

Control Palos Verdes, CA

Young et al., 1978

Highly contaminated sediments,

Species H.multirugosus

0.41±0.1 mg/kg

Adductor muscle

3.2±0.1 mg/kg

Gonad

Level

Comments

References

190±40 mg/kg

Digestive gland Natural,

0.16±0.04 mg/kg

Adductor muscle

2.2±0.5 mg/kg

Gonad

64±15 mg/kg

Digestive gland

0.24 mg/kg

Palos Verdes, CA

Young & Mearns, 1978

Ischadium demissium

av. 20 mg/kg

San Francisco, U.S.A.

Anderlini et al., 1975

Lingula unguis

7.00–12.18

Ariake Sea

Enomoto & Toyodome, 1978

Macoma balthica 49.1–313.6 96–615 (m. 300)

Young et al., 1978

Fraser River estuary, B.C., McGreer, 1979b Canada Looe estuary, U.K.

Bryan & Hummerstone, 1977

San Francisco, U.S.A.

Anderlini et al., 1975

30 mg/kg

Before dredging

90 mg/kg

After dredging

88.1–171

St. Croix River, Maine

Fink et al., 1976

M.modialus

av. 27 mg/kg

San Francisco, U.S.A.

Anderlini et al., 1975

Mercenaria mercenaria

15.8–35.4 mg/kg freeze dry wt

Thames River estuary, Conn.

National Marines Fisheries Service, 1976b

Interlinking of physical

Meretrix chionae

1146

25

Irish Sea

Segar et al., 1971

1.0–16.50

East coast, U.S.A.

Pringle et al., 1968

0.6–7.9 mg/kg wet wt Lower Chesapeake Bay (range of m. 1.0–3.7)

Larsen, 1979

84 mg/kg ash wt

Saronikos Gulf, Greece

Papadopoulou & Kanias, 1976

52.7 mg/kg ash wt

Muscle

64.8 mg/kg ash wt

Mantle and gills

153 mg/kg ash wt

Stomach and intestines

114 mg/kg ash wt

Gonads

136 mg/kg ash wt

Foot

M.lamarckii

15

Japan

Ishii et al., 1978

Modiolus modiolus

44

Irish Sea

Segar et al., 1971

Mya arenaria

8.4–21.3

St. Croix River, Maine

Fink et al., 1976

23 mg/kg

Atlantic coast, U.S.A.

Anderlini et al., 1975

1.20–90 (m. 5.80)

East coast, U.S.A.

Pringle et al., 1968

28

Cowichan Bay, Canada

Bawden et al., 1973

Mytilaster lineatis 5.8

Sea of Azov

Rozhanskaya, 1967

Mytilis californianus

9

California

Graham, 1972

30.3

”

”

10.7

”

”

av. 25

Outer Islands S.California Bight

Alexander & Young, 1976

1.7–10.9 mg/kg

S.California Bight, soft parts

Callahan & Shokes, 1978

6.5–12

S.W.Atlantic coast, Spain—Portugal (6 locations)

Stenner & Nickless, 1975

1.7–18.0 mg/kg wet wt (m. 8.3)

New Zealand

Nielsen & Nathan, 1975

m. 0.4–1.45

Port Phillip Bay, Australia Phillips, 1976a (23 locations)

3.9–13.6 (m. 9.5)

Looe estuary, U.K.

(m. 5.0) M.edulis

Bryan & Hummerstone, 1977

Oceanography and marine biology

673

0.5–3.8 mg/kg wet wt Medway estuary, U.K.

Species

Level

M.edulis 7–11 mg/kg

Wharfe & Van Den Broek, 1977

Comments

References

Poole Harbour, U.K.

Boyden, 1975

5.5–9.0 mg/kg

Tamales Bay, U.S.A.

Anderlini et al., 1975

5–88

Norwegian fjord

Lande, 1977

4.8

Shells

3–22

Hardangerfjord, Norway

Stenner & Nickless, 1974a

15–130

Skjerstadtfjord, Norway

”

2.4–52.9

Vinogradov, 1953*

5.3–11.2

California

Graham, 1972

9

New Zealand

Brooks & Rumsby, 1965

9.6

Irish Sea

Segar et al., 1971

3.30–5.61 mg/kg wet wt

Iona Island, B.C., Canada

Vermeer & Peakall, 1979

2.33 mg/kg wet wt

Roberts Bank, B.C., Canada Southern B.C.

5–898

Gills

10–1506

Viscera Newport, CA (harbour)

127±18

Digestive gland

93±15

Gonad

52±10

Muscle

100±14

Remainder

16±1.0

Digestive gland

9.6±0.2

Gonad

5.7±0.3

Muscle

11±0.6

Remainder

Newport, CA beach

Royal Palms, CA 47±2.8

Digestive gland

42±0.5

Gonad

14±1.0

Muscle

44±7.0

Remainder

Stump et al., 1979

Young & Alexander, 1977

”

”

Interlinking of physical

1148

Other coast stations, CA 20±2.4

Digestive gland

10±0.8

Gonad

11±2.9

Muscle

13±1.5

Remainder

11.5

Knokke, Belgium

10.77

Gills

7.2

Foot

5.85

Mantle

8.72

Kidney

6.87

Digestive gland

2.74

Muscles

8.37

Morgat, Brittany

6.91

Gills

4.21

Foot

5.4

Mantle

7.4

Kidney

7.02

Digestive gland

3.72

Muscles Whole, Elkhorn Slough, CA

30±6.7 µg/g

Redwood City, CA

M.edulis

Level

Delhaye & Cornet, 1975

”

Martin, 1979

19±6 µg/g

Species

”

Comments

18.8

Moss Landing, CA unpolluted

30.4

San Francisco, CA polluted

References Martin, 1979

0.3–0.5 mg/kg Gulf of Thessaloniki, Greece

Kovatsis et al., 1975

4.5–5.6 mg/kg S.California Bight, soft parts

Callahan & Shokes, 1978

(m. 5.1) 14.6–14.9

Restronguet Creek

Klumpp & Peterson, 1979

Western Port, Australia (22 locations)

Harris et al., 1979

(m. 14.8) M.edulis planulatus

m. 1.7–8.2

M.galloprovincialis 45.1 mg/kg ash Saronikos Gulf, Greece

Papadopoulou & Kanias,

Oceanography and marine biology

675

wt

1976

14.6 mg/kg ash Muscle wt 50.1 mg/kg ash Mantle and gills wt 83 mg/kg ash wt

Stomach and intestines

26.5 mg/kg ash Gonads wt 36.3 mg/kg ash Foot wt 12.8 mg/kg ash Byssus wt

Mytilus sp.

6.9–33.7

Port of La Spezia, Italy

Capelli et al., 1978

3.3–4.2

Promontory of Portafino, Italy

”

8.35–15.0

E.Sicillian coast, Italy

Castagna & Sarro, 1975

2.4–154

N.W.Mediterranean Sea

Fowler & Oregioni, 1976

6.9

California

Graham, 1972

7–15

Dutch coast

Hueck, 1975

5.63–6.86

S.California Bight

Science Applications Inc., 1978c

San Francisco, U.S.A.

Anderlini et al., 1975

Neanthes succonea 25 mg/kg

before dredging

45 mg/kg

after dredging

Neanthes sp.

31–61

Texas, 30‰ salinity

Neff et al., 1978

Ostrea edulis

1609 mg/kg ash wt

Saronikos Gulf, Greece

Papadopoulou & Kanias, 1976

368 mg/kg ash Muscle wt 590 mg/kg ash Mantle and gills wt 518 mg/kg ash Stomach and intestines wt 370 mg/kg ash Gonads wt 398 mg/kg ash Foot wt

Interlinking of physical

1150

112.6–259.0

Cadiz coast, Spain

Establier & Gutierrez, 1970

86–451

Poole Harbour, U.K.

Boyden, 1975

23 mg/kg wet wt

Anglesey, U.K.

George et al., 1978

207 mg/kg wet Mantle wt 64 mg/kg wet wt

Muscle

253 mg/kg wet Gills wt 124 mg/kg wet Kidney wt 77 mg/kg wet wt

Viscera

2.0 mg/kg wet Haemolymph wt 221 mg/kg wet Cornwall, U.K. wt 2175 mg/kg wet wt

”

Mantle

315 mg/kg wet Muscle wt 2395 mg/kg wet wt

Gills

500 mg/kg wet Kidney wt 1184 mg/kg wet wt

Species Ostrea edulis

Viscera

Level

Comments

References

21 mg/kg wet wt

Haemolymph

George et al., 1978

2132–2220

Restronguet Creek

Klumpp & Peterson, 1979

Ariake Sea

Enomoto & Toyodome, 1978

(m. 2161) Ostrea sp.

111.59–114.57

O.lurida

4–12.4 mg/kg wet wt

O.lutaria

1.0–41.0 mg/kg wet

Vinogradov, 1953* New Zealand

Nielsen & Nathan, 1975

Oceanography and marine biology

677

wt (m. 11.0) O.sinuata

21–53

”

Brooks & Rumsby, 1965

Green oysters

1180 mg/kg

Kaohsinmg Bay, Taiwan

Chang, 1965

Normal oysters

5.6 mg/kg

Taiwan

Patella vulgata

7.7

Irish Sea

Pecten circularis laquisulcatus

24.0

Vinogradov, 1953*

P.fumatus

0.1 mg/kg wet wt

”

P.jacobaeus

14–47 mg/kg wet wt

”

P.maximus

m., 8.9, SD, 4.5

English Channel

Bryan, 1973

3.3

Irish Sea

Segar et al., 1971

1.1–2.3 mg/kg wet wt P.novae-zeolandiae

Segar et al., 1971

2–14

Vinogradov, 1953* New Zealand

Brooks & Rumsby, 1965

m. 1.46 mg/kg wet wt New Zealand

Nielsen & Nathan, 1975

m. 6.6 mg/kg wet wt

Muscle

2.1 mg/kg wet wt

Gonad

P.varius

3.4 mg/kg wet wt

Vinogradov, 1953*

Pectunculus glycemaris

4.93 mg/kg wet wt

”

Perna canaliculus

0.2–28.0 mg/kg wet wt (m. 1.8)

Pitar morrhuana

11.1–19 mg/kg freeze Thames River dry wt estuary, Conn.

National Marine Fisheries Service, 1976b

2.3–29.6 mg/kg

Rhode Island, Soft parts

Eisler et al., 1977a

4.01–12.65 mg/kg wet wt (m. 7.31)

U.S.A. Mid-Atlantic Pesch et al., 1977 coast

Placopecten magellanicus

New Zealand

Nielsen & Nathan, 1975

Mid-Atlantic, U.S.A.

Protothaca stamina

Rangia cuneata Rangia sp.

m. 0.80–14.84

Muscle

Reynolds, 1979

m. 6.99–20.09

Viscera

8 mg/kg

San Francisco, U.S.A.

m. 6.1

Puget Sound, U.S.A. Stober et al., 1977

22.75 mg/kg wet wt

Galveston, Texas

Guthrie et al., 1979

Texas

Neff et al., 1978

Anderlini et al., 1975

Interlinking of physical

1152

12–16

0‰ salinity

14–22

15‰ salinity

8.7

30‰ salinity

Scrobicularia plana

1 6–365 (m. 133)

Looe estuary, U.K.

Scrobicularia sp.

13–52(m. 34)

Tamar estuary, U.K. Bryan & Uysal, 1978

26–102 (m. 58)

Digestive gland

25–86

Gannel estuary, U.K. Bryan & Hummerstone, 1978

92

Digestive gland

29

Mantle and siphon

48

Foot and gonad

36

Gills and palps

5.3

Adductor muscles

Species Scrobicularia sp.

Level

Comments

Bryan & Hummerstone, 1977

References

37

Kidney

25–77

Camel estuary, U.K.

51

Digestive gland

30

Mantle and siphon

19

Foot and gonad

45

Gills and palps

8.3

Adductor muscles

28

Kidney

Sinonovacula constricta

6.99–44.54

Ariake Sea

Enomoto & Toyodome, 1978

Spisula solidissima

m. 2.87–3.83

Mid-Atlantic coast, U.S.A. range of means

Wenzloff et al., 1979

Spondylus americanus

3.1–87

Texas

Rezak et al., 1978

Tapes decussatus

103 mg/kg ash wt Saronikos Gulf, Greece

Papadopoulou & Kanias, 1976

2.26–50.7 mg/kg Spain wet wt

Establier, 1978

Azmur Bay area, Turkey,

Bryan & Hummerstone, 1978

Geldiay & Uysal, 1976

Oceanography and marine biology

679

5.56

Blood

7.07

Foot and digestive system

10.40

Mantle and siphon

1.74

Muscle

7.80

Gills and palps

3.19

Digestive gland

9.82

Soft body

T.japonica

19 mg/kg

San Francisco, U.S.A.

Anderlini et al., 1975

Venerupis philippinarum

3.83–12.24

Ariake Sea

Enomoto & Toyodome, 1978

Venus gallina

63.10–156.95

Spain

Establier, 1978

V.verrucosa

73.8 mg/kg ash wt

Saronikos Gulf, Greece,

Papadopoulou & Kanias, 1976

78.1 mg/kg ash wt

Muscle

85.4 mg/kg ash wt

Mantle and gills

51.7 mg/kg ash wt

Stomach and intestines

58.7 mg/kg ash wt

Gonads

2498 mg/kg ash wt

Foot

Amphineura Cryptochiton stelleria

2.5 mg/kg wet wt

Vinogradov, 1953*

Ischnochiton conspicuus

2.5–3.7 mg/kg wet wt

”

Cephalopoda Alloteuthis subulata m. 5.78, SD, 2.38 From Belgian inshore Fishery De Clerck et al., 1979 mg/kg Loligo opalescens

15, 160

L.vulgaris

16

Octopus vulgaris

47

Octopus 3.39 mg/kg ash

California, liver

Martin, 1975 Vinogradov, 1953*

S.W. Atlantic coast, Spain— Portugal

Stenner & Nickless, 1975

Japan

Matsumoto et al., 1965

Whole

Interlinking of physical

1154

wt 0.30 mg/kg ash wt

Muscle

0.44 mg/g ash wt Eyeball 1.40 mg/g ash wt Ink 2.15 mg/g ash wt Gill 2.26 mg/g ash wt Heart 11.66 mg/g ash wt

Liver

0.57 mg/g ash wt Gonad and digestive organ

Species Octopus

Omnastrephes bartrami

Level 1.53 mg/g ash wt

Comments

References

Excretory organ

Matsumoto et al., 1965

San Miguel Island, CA

Callahan & Shokes, 1978

1220+180

Digestive gland

271+20

Body

548+24

Santa Rosa Island, CA

17–696

Pacific Ocean, liver

Martin, 1975

Fresh,

Schipp & Hevert, 1978

Sepia officinalis m. 157, SD 27mg/kg

Blood

m. 100, SD 49 mg/kg

Hepatopancreas

m. 22, SD 7 mg/kg Branchial gland m. 20, SD 7 mg/kg Branchial heart m. 28, SD 8 mg/kg Branchial heart appendage m. 23, SD 4 mg/kg Heart ventricle m. 20, SD 7 mg/kg Pancreatic appendages Wet, m. 1182, SD 277 mg/kg

Blood

m. 298, SD 131 mg/kg

Hepatopancreas

m. 122, SD 43

Branchial gland

Oceanography and marine biology

681

mg/kg m. 111, SD 28 mg/kg

Branchial heart

m. 219, SD 76 mg/kg

Branchial heart appendage

m. 112, SD 26 mg/kg

Heart ventricle

m. 93, SD 38 mg/kg

Pancreatic appendages

Sepia

Japan 1.78 mg/g ash wt

Whole, male

1.81 mg/g ash wt

Whole, female

0.24 mg/g ash wt

Muscle

0.11 mg/g ash wt

Eyeball

0.26 mg/g ash wt

Ink

1.61 mg/g ash wt

Gill

Matsumoto et al., 1965

17.57 mg/g ash wt Liver 0.54 mg/kg ash wt Gonad and digestive organ, male

Sepioteuthis indica

0.79 mg/g

Gonad and digestive organ, female

2.30–2.96

Strait of Singapore

Bok & Keong, 1976

Strait of Malacca

”

Pacific Ocean, livers

Martin, 1975

(m. 2.49) 1.71–2.58 (m. 2.20) Symplectoteuthis oualaniensis

m. 1720

Gastropoda Gastropods

1–554

Reish et al., 1978*

Acmaea digitalis

11.2–12.9

California

Graham, 1972

A.scabra

m. 3.3–8.1

California coast (12 locations)

Flegal, 1978

5.6

S.California Bight, Soft parts

Callahan & Shokes, 1978

Amphibola crenata

48.9–262.6

New Zealand

Stephenson, 1978

Aplysia punctata

7 mg/kg wet wt

Plymouth, U.K.

McCance & Shackleton, 1937

Interlinking of physical

1156

Archidoris montereyensis

1.3 mg/kg wet wt

Vinogradov, 1953*

A.undosa

8 mg/kg wet wt

”

A.britannica

24 mg/kg wet wt

”

Species

Level

Comments

References

Bathybembrix sp. 76.6±34– 125±107

Outer Continental Shelf, >200m

Callahan & Shokes, 1978

Buccinum undatum

180

Irish Sea

Segar et al., 1971

Busycon canaliculatum

58–116 µg/g wet wt

Rhode Island,

Betzer & Pilson, 1974

65–900 µg/g

Digestive gland

Caliostoma zizyphinium

Plymouth U.K.,

McCance & Shackleton, 1937

54 µg/g wet wt Foot and gut 110 µg/g wet wt

Gonad and liver

Chione undatella 1.2–2.0 mg/kg wet wt

Vinogradov, 1953*

C.stutchburyi

9.6–371.8

New Zealand

Stephenson, 1978

Cominella adspersa

36.6–117.2

”

”

Cypraea spadacea

51.7–779 mg/kg

S.California Bight, soft parts

Callahan & Shokes, 1978

Irish Sea

Segar et al., 1971

(m. 316) Crepidula fornicata

270

Donax gouldii

0.6–0.7 mg/kg wet wt

Vinogradov, 1953*

Haliotis cracherodii

0.8 mg/kg wet wt

”

3.4 mg/kg wet wt

S.California wastewater outfall, JWPCP

3.9 mg/kg wet wt

Island control

2.7–128 mg/kg S. California Bight, Digestive gland

Jan et al., 1977

Callahan & Shokes, 1978

Oceanography and marine biology

683

(m. 21.4) H.rufescens

California coast samples taken from 5 locations N→S along coast 18.8–123.5

Gill

8.9–23.1

Mantle

1.4–12.8

Foot

11.2–78.1

Digestive gland

Anderlini, 1974

H.striata

40 mg/kg wet wt

Vinogradov, 1953*

H.tuberculata

1.2 mg/kg wet wt

”

Haliotis sp.

3.69

Helix aspersa

9.28 mg/kg wet wt

Vinogradov, 1953*

Joruna tormentosa

21.6 mg/kg wet Plymouth, U.K. wt

McCance & Shackleton, 1937

S.California Bight

Littorina littoralis 47.3 mg/kg wet Plymouth, U.K., Foot and gut wt

Science Applications Inc., 1978c

”

20–170

Norwegian fjords

Stenner & Nickless, 1974a

m. 42.7–248.8 mg/kg

9 locations on Welsh coast

Ireland & Wooton, 1977

19.5–40.0 mg/kg wet wt

Lower Medway estuary, U.K.

Wharfe & Van Den Broek, 1977

62–194

Looe estuary, U.K.

Bryan & Hummerstone, 1977

(m. 124) 142 mg/kg wet Restronguet Creek, Cornwall wt (contaminated)

Howard & Nickless, 1978

14.4 mg/kg wet Brighton, Sussex, Natural (relatively ” wt uncontaminated) 46.6 mg/kg wet Sand Point, Somerset (contaminated) ” wt 54.6–68.0 L.neritoides

St. Croix River, Maine

102 mg/kg wet Plymouth, U.K. wt

Fink et al., 1976 McCance & Shackleton, 1937

Interlinking of physical

Species L.obtusa

Level

1158

Comments

4.3 mg/kg wet wt

L.rudis

References Vinogradov, 1953*

Plymouth, U.K., 31 mg/kg wet wt

Foot and gut

81 mg/kg wet wt

Gonad and liver

L.saxatilis

80

Norwegian fjord

Lottia gigantea

1.3–6.8 mg/kg S.California Bight, Soft parts

McCance & Shackleton, 1937

Stenner & Nickless, 1974a Callahan & Shokes, 1978

(m. 2.9) Monodonta turbinata

45.0–135.0

Ras Beirut, Lebanon

Shiber & Shatila, 1978

(m. 87.1) Murex tronchulus 21.1 mg/kg wet wt

Vinogradov, 1953*

Navanax merimis

”

4.4 mg/kg wet wt

Norrissia norrissii 40.3–82.8 mg/kg

S.California Bight

Callahan & Shokes, 1978

Plymouth, U.K.

McCance & Shackleton, 1937

(m. 61.6) Nucella lapillus 15 mg/kg wet wt

Foot and gut

53 mg/kg wet wt

Gonad and liver

20–110

Dorset, U.K.

385–1750

Somerset, U.K.

150

Stenner & Nickless, 1974b

Irish Sea

Segar et al., 1971

Hinkley Power Station,

Boyden & Romeril, 1974

282

U.K., outfall

250

Intake

Oceanography and marine biology

51–141

685

Looe estuary, U.K.

Bryan & Hummerstone, 1977

Singapore coast

Bok & Keong, 1976

(m. 110) Paphia luzonica

1.27–3.17 mg/kg wet wt

P.staminea var. laciniata

1.8–2.4 mg/kg wet wt

Paphia sp.

8.4–20.2 mg/kg ash wt

Hong Kong

15.5–30.0 mg/kg ash wt

Shell

Vinogradov, 1953* Wong & Li, 1977

Patella athletica

6.7 mg/kg wet Plymouth, U.K. wt

McCance & Shackleton, 1937

P.coerulea

11.3–38.0 (m. Ras Beirut, Lebanon 27.8)

Shiber & Shatila, 1978

P.intermedia

1.16 mg/kg wet wt

Newquay, England

Howard & Nickless, 1977a

P.vulgaris (sic)

17

Norwegian fjords

Stenner & Nickless, 1974a

P.vulgata

9.7 mg/kg wet Plymouth, U.K. wt

McCance & Shackleton, 1937

28.5 & 45.5

Hinkley Power Station, U.K., settling Boyden & Romeril, tank, Jan. 1 and Mar. 2, 1973 1974

20.2

Foreshore

27.4

Outfall

34.0

Intake Sewage discharge pipe Tel-Aviv

Navrot et al., 1974

Soft parts: 10.5

Discharge site

11.2

North of site

5.5–8.2

Other areas Skeletal parts:

5.2

Discharge site

4.9

North of site

4.6–6.2

Other areas

12–22

Norwegian fjords,

Lande, 1977

Interlinking of physical

Species P.vulgata

Level

1160

Comments

References

1–6

Shells

Lande, 1977

3.83 mg/kg wet wt

Aberystwyth, U.K.

Howard & Nickless, 1977a

3.80 mg/kg wet wt

Sand Point, U.K.

”

11.9 mg/kg wet wt

Mylor, U.K.

”

7.4 mg/kg wet wt

Brean Down, U.K.

Howard & Nickless, 1975

10–27 (m. 19)

Looe estuary, U.K.

Bryan & Hummerstone, 1977

9–282 mg/g dry wt (sic)

Marine–estuarine

Leland et al., 1978*

42.3–45.9

Restronguet Creek, U.K. Klumpp & Peterson, 1979

(m. 44.7) Patella sp.

5.23–19.77

Polypus bimaculatus

31.5 mg/kg wet wt

Protothaca staminea

m. 7.5

California, body

m. 11.5

Shell

Saxostrea commercialis

0.3 mg/kg wet wt

Scaphander lignarius

5 mg/kg wet wt

Strombus gigas

Tegula funebralis

Spain

Establier, 1978 Vinogradov, 1953* Graham, 1972

Vinogradov, 1953* Plymouth, U.K.

McCance & Shackleton, 1937

Biscayne Bay, Fla.

Oppenheimer, 1968

1330 mg/kg ash wt

Liver

600 mg/kg ash wt

Kidney

170 mg/kg ash wt

Muscle

m. 14–175

California, body (5 stations)

m. 6.8–19.9

Shell (6 stations)

14.6–148 mg/kg S.California Bight, Soft parts

Graham, 1972

Callahan & Shokes, 1978

Oceanography and marine biology

687

(m. 65.4) T.galina

6.1 mg/kg wet wt

Vinogradov, 1953*

T.viridula

3.1–6.1 mg/kg wet wt

”

Thais emarginata

m. 571.0

California, body

m. 5.9–8.2

Shell

27–170

Norwegian fjords

Stenner & Nickless, 1974a

m. 116.4–458.1 mg/kg

9 locations on Welsh coast

Ireland & Wooton, 1977

T.lapillus

Tivela crassatelloides

2.1–2.3 mg/kg wet wt

Turbo cornutus

Graham, 1972

Vinogradov, 1953* Ariake Sea

Enomoto & Toyodome, 1978

2.17–3.24

Flesh

2.71–6.63

Internals

79.6–166.8

New Zealand

Crustaceans

1.0–20.0 mg/kg

198 sites, coastal U.S.A. Hall et al., 1978

Decapods

2–175

Reish et al., 1978*

6–63.7 µg/g wet England wt

Reish et al., 1977*

T.smaragdus

Stephenson, 1978

Crustacea

Blue crab

Brown shrimp

Crab

54

San Antonio Bay, Texas Sims & Presley, 1976

34

U.S.A., S.E.coast

34

San Antonio Bay, Texas ”

24

N.E.Mexican coast

”

11.5

Mississippi River

”

10.34 mg/kg wet Mediterranean, leg wt

Lobsters

New London dump site, U.S.A., 370–775 mg/kg wet wt

Species Lobsters

Level

”

Ramelow et al., 1978 National Marine Fisheries Service, 1976a

Digestive diverticulata

Comments

4.8–72 mg/kg Tail muscle wet wt

References National Marine Fisheries Service, 1976a

Interlinking of physical

19–42 mg/kg wet wt

1162

Gills New Haven dump site, U.S.A., ”

480–2308 Digestive diverticulata mg/kg wet wt 5.7–17.5 Tail muscle mg/kg wet wt Prawn

41.2 & 12.7 µg/g

Off N.W.Africa

Brugmann, 1978.

Shrimp

1.77 mg/kg wet wt

Mediterranean

Ramelow et al., 1978

White shrimp

17

San Antonio Bay, Texas

Sims & Presley, 1976

Santa Barbara Channel,

McDermott-Ehrlich & Alexander, 1976

Yellow rock crab 33–41

Muscle

17–45

Gonad

10

Plymouth, U.K.

Bryan, 1968

Lagoon of Venice

Barbare et al., 1978

4–54

Lagoon of Grado, Italy

”

500–550

S.W.Atlantic coast, Spain– Portugal (2 locations)

Stenner & Nickless, 1975

B.aburneus (sic)

15.28 mg/kg wet wt

Galveston, Texas

Guthrie et al., 1979

B.balanoides

3.75

Dulas Bay, U.K.

Walker, 1977

B.improvisus

8.6

Cornwall, U.K.

Rozhanskaya, 1967

B.perforatus

8.0–8.5

S.W.Atlantic coast, Spain– Portugal (2 locations)

Stenner & Nickless, 1975

Callianassa californianus

330–4100

British Columbia, Canada

Bawden et al., 1973

Georgia estuaries, U.S.A.

Stickney et al., 1973

Atelecyclus septemdentatus

Balanus amphithrite 96–109

Calllnectes sapidus

Cancer anthonyi

31

Muscle

41

Egg mass

60

Gills

2.99 mg/kg wet wt

Galveston, Texas

7.9 mg/kg wet S.Calif, waste water outflow

Guthrie et al., 1979 Jan et al., 1977

Oceanography and marine biology

689

wt 13 mg/kg wet Coastal control wt 7.84 mg/kg

Palos Verdes, CA

Young & Mearns, 1978

C.magister

22–150

British Columbia, Canada

Bawden et al., 1973

C.pagurus

6.80 mg/kg wet wt

Northumberland coast, U.K.

Wright, 1976

20

Plymouth, U.K.

Bryan, 1968

C.productus

18–57

British Columbia, Canada

Bawden et al., 1973

Carcinus maenas

15–175

Estuarine

Leland et al., 1978*

Species Carcinus maenas

15.4–31.4 Lower Medway estuary U.K. mg/kg wet wt

Wharfe & Van Den Broek, 1977

175

Orkdalsfjorden, Norway

Lande, 1977

1.7–90

Norwegian fjords

Stenner & Nickless, 1974a

Northumberland coast, U.K.

Wright, 1976

8.70 mg/kg wet wt

Exoskeleton (carapace)

10.90 mg/kg wet wt

Gills

6.00 mg/kg wet wt

Muscle

Level

Comments

References

22.90 mg/kg wet wt

Hepatopancreas

Wright, 1976

10.90 mg/kg wet wt

Whole body

63.70 mg/kg wet wt

Haemolymph

22–40.6

Plymouth, U.K.

Bryan, 1968

148.9–276.1

Restronguet Creek, U.K.

Klumpp & Peterson, 1979

(m. 221.2) Chthamalus stellatus

6.3–11.3

S.W.Atlantic coast, SpainPortugal (3 locations)

Stenner & Nickless, 1975

Crangon allmani

56–112

Firth of Clyde, U.K.

Halcrow et al., 1973

C.vulgaris

1.3–18.2 µg/g wet Medway estuary, U.K.

Van Den Broek,

Interlinking of physical

1164

wt

1979

9.5–25.5 mg/kg. wet wt

Lower Medway estuary, U.K. Wharfe & Van Den Broek, 1977

19 mg/kg dry wt

Marine

49

Hinkley, England, Tail muscle Boyden & Romeril, 1974 Northumberland coast, U.K.,

Leland et al., 1978*

Wright, 1976

7.00 mg/kg wet wt Exoskeleton (carapace) 25.9 mg/kg wet wt Muscle 32

Plymouth, U.K.

Bryan, 1968

East Anglesey, U.K.

Djangmah & Grove, 1970

41–159 µg/ml

Blood

13–480

Hepatopancreas West Anglesey, U.K.

54–200 µg/ml

Blood

118–2900

Hepatopancreas

Emerita analoga

S.California

Callahan & Shokes, 1978

81±21 µg/g

Mainland sites

50±17 µg/g

1st line island sites

85±37 µg/g

2nd line island sites

26.5–136 mg/kg

S.California Bight, whole

”

(m. 73.2) Emerita sp.

64.2–92.1

S.California Bight

Science Applications Inc., 1978c

Eupagurus bernhardus

25

Plymouth, U.K.

Bryan, 1968

Eupagurus sp.

20

”

”

Galathea squamifera

29

”

”

Hemigrapsus sp.

59.1–62.5 mg/kg wet wt

Iona Island, B.C., Canada

Vermeer & Peakall, 1979

Georgia estuaries.

Stickney et al., 1975

Hexapanopeus augustifrons 13

U.S.A., muscle

Oceanography and marine biology

691

Homarus vulgaris

33

Plymouth, U.K.

Bryan, 1968

Leander adspersus

39–60 mg/kg ash wt

Black Sea

Petkevich & Stepanyuk, 1970

L.squilla

40–41 mg/kg ash wt

”

”

Maia squinado

25

Plymouth, U.K.

Bryan, 1968

Meganyctiphanes norvegicn

209.97 mg/kg

Strait of Messina, Italy

Calapaj et al., 1978

Strait of Singapore,

Bok & Keong, 1976

2.09–3.36 mg/kg wet wt (m. 2.52)

Flesh

Neptunus pelagicus

Species Neptunus pelagicus

Pachygrapsus crassipes

Level

Comments

References

6.08–8.42 mg/kg wet wt (m. 7.20)

Whole

Bok & Keong, 1976

3.90–4.85 mg/kg wet wt (m. 3.90– 4.85)

Strait of Malacca

”

123–284 mg/kg

S.California Bight, Digestive gland

Callahan & Shokes. 1978

Coast of Ras Beirut, Lebanon

Shiber, 1979

(m. 182) Palaemon elegans

P.serratus

112.5–176.3

Summer 1977

76.9–95.6

Fall, 1977

111.6–216.6

Spring, 1978

30

Plymouth, U.K.

25 mg/kg wet wt

Bryan, 1968 Vinogradov, 1953*

P.squilla

31

Plymouth, U.K.

Bryan, 1968

Palaemonetes pugio

74

Georgia estuaries, U.S.A.

Stickney et al., 1975

Whole animal P.varians

32

Palaemonetes spp. (P.kadiakensis and P.pugio)

Plymouth, U.K.

Bryan, 1968

Texas

Neff et al., 1978

54–91

0‰ salinity

47–130

15‰ salinity

Interlinking of physical

Pandalus montagui

1166

93–181

30‰ salinity

77–113

Firth of Clyde, U.K.

Panulirus interrupts

Halcrow et al., 1973

S.California waste water Jan et al., 1977 outflow, JWPCP 6.1 mg/kg wet wt 6.4 mg/kg wet wt

Coastal control

14 mg/kg wet wt

Island control

Penaeus aztecus

18

Georgia estuaries, U.S.A. Muscle

Stickney et al., 1975

P.brasiliensis

20–167.50 µg/g

Caribbean, N.E. of Margarita, Venezuela

Morales & Shrestha, 1977

Strait of Singapore,

Bok & Keong, 1976

P.indicus 3.37–5.76 mg/kg wet wt (m. 4.70)

Flesh

14.4–15.1 mg/kg wet wt (m. 14.8)

Whole

3.13–3.34 mg/kg wet wt (m. 3.20)

Strait of Malacca

”

14

Norway

Lande, 1977

38

Hinkley, England, Tail muscle

Boyden & Romeril, 1974

P.semisulcatus

83.2–160

Black Sea

Petkevich & Stepanyuk, 1970

P.setiferus

16

Georgia estuaries, U.S.A., muscle

Stickney et al., 1975

Pilumnus hirtellus

28

Plymouth, U.K.

Bryan, 1968

Pollicipes polymerus

1.6–3.7 mg/kg

S.California Bight, Muscle

Callahan & Shokes, 1978

P.japonica

(m. 2.4) Porcellana platycheles

27

Plymouth, U.K.

Bryan, 1968

Portunus depurator

18

„

”

P.holsatus

m. 12.23, SD 3.83 mg/kg

From Belgian inshore Fisheries

De Clerck et al., 1979

P.puber

21

Plymouth, U.K.

Bryan, 1968

Scopimera intermedia

105–252.5

Hong Kong

M.H.Wang et al., 1978

Oceanography and marine biology

Sicyonia ingentis

2.0 mg/kg wet wt

Species Sicyonia ingentis

Level

693

S.California waste water Jan et al., 1977 outflow, JWPCP

Comments

References

8.0 mg/kg wet wt Orange County

Jan et al., 1977

2.0 mg/kg

Palos Verdes, U.S.A.

Young & Mearns, 1978

Squilla oratorio

10.37–68.80

Ariake Sea

Enomoto & Toyodome, 1978

Trachypencus constrictus

56

Georgia estuaries, U.S.A., muscle

Stickney et al., 1975

Tylos punctatus

m. 13–59

San Diego, whole

Hayes, 1970

m. 200

Hepatopancreas

79–190

British Columbia, Canada

Bawden et al., 1973

Georgia estuaries, U.S.A., muscle

Stickney et al., 1975

Upogebia pugettensis

Xiphopeneus kroyeri 22 Echinodermata Echinoids

0.26–0.27 µg/g wet wt

California

Reish et al., 1978*

Allocentrotus sp.

30.6±4.6

Outer continental shelf ≥200 m depth

Callahan & Shokes, 1978

1.8±0.6

Tanner Cortes Banks, S.California

5.1 ±0.8

Outer continental shelf ≥200 m depth

2.7±0.9

Various island locations

7.0

Park, Italy

16.6

Bay of Naples, Italy

4.3

Irish Sea

Riley & Segar, 1970

18

Norwegian fjords

Stenner & Nickless, 1974a

Arbacia lixula

Asterias rubens

Sheppard & Belamy, 1974

m. 3.59, SD 0.87 From Belgian inshore mg/kg Fisheries

De Clerck et al., 1979

Asteroideae

0.0265 mg/g wet wt

Mid-Atlantic Bight

Palmer & Lear, 1973

Brissopsis sp.

24.5 ±13.9

Tanner-Cortes Banks,

Callahan & Shokes,

Interlinking of physical

1168

S.California Echinarachnius

0.04–11.4 mg/kg Mid-Atlantic Bight

Echinus esculentus

Irish Sea,

1978 Palmer & Lear, 1973 Riley & Segar, 1970

1.8

Oral shell

0.9

Aboral shell

0.42

Aristotle’s lantern

1.6

Spines

5.9

Intestine

16.0

Gonad

Henrica sanguinolenta

8.2

Irish Sea

”

Holothuria

0.06 mg/g ash

Japan

Matsumoto et al., 1965

Luidia sp.

5.2+0.9 & 5.4 ±2.7

Various island locations, Continental Shelf

Callahan & Shokes, 1978

Paracentrotus lividus

7.6

Park, Italy

Sheppard & Belamy, 1974

13.8

Bay of Naples, Italy Coast of Ras Beirut, Lebanon

Shiber, 1979

11.3–38.0

Summer 1977

3.8–<744.4

Fall 1977

11.3–28.1

Spring 1978

Patiria miniata

16.7–36 mg/kg (m. 26.5)

S.California Bight, Digestive gland

Callahan & Shokes, 1978

Patiria sp.

19.8–29.3

S.California Bight

Science Applications Inc., 1978c

Pisaster giganteus

6.7–31.5 mg/kg (m. 17.9)

S.California Bight, Digestive gland

Callahan & Shokes, 1978

Species P.ochraceus

Level 8.8–62.3 mg/kg

Comments

References

S.California Bight, Digestive gland

Callahan & Shokes, 1978

(m. 32.0) Pisaster sp.

29.4

S.California Bight

Science Applications Inc., 1978c

Porania pulvillus

30

Irish Sea

Riley & Segar, 1970

Solaster papposus

6.1–11

”

”

Oceanography and marine biology

Strongylocentrotus droebachiensis

S.franciscanus

695

Strathcona Sound, Canada Pre- Bohn, 1979 mining of Pb-Zn ore, 2.3–9.0

Whole

3.3–9.0

Gonads

0.27 mg/kg wet wt

S.California waste water outflow, JWPCP

0.26 mg/kg wet wt

Island control

1.0–4.6 mg/kg

S.California Bight Internal organs

Callahan & Shokes, 1978

S.California Bight Internal organs

”

S.Africa

Reish et al., 1978*

Jan et al., 1977

(m. 2.4) S.purpuratus

2.8–9.5 mg/kg (m. 5.2)

Chordata Tunicates

0.3–0.78 µg/g wet wt

Ciona intestinalis

55.12–205.40 Gulf of Napoli, Neopolitan Harbour

Melluso et al., 1978

decrease with age and weight (on a weight: weight basis) although total metal concentrations increase in Crassostrea commercialis. Copper concentrations in Mercenaria mercenaria from the lower Chesapeake Bay region varied significantly with the age of the organism (Larsen, 1979). Harris et al. (1979), as well as others, have found the metal and water content of mussels to be linear functions of length. Copper concentrations in the “mediterranean mussel” (Ramelow et al., 1978) and slope coefficients for 12 species of marine or estuarine molluscs examined by Boyden (1977) indicate a general increase in copper with size. Seasonal fluctuations of trace metal levels in Crassostrea gigas along the Washington coast have been documented by Pringle, Hissong, Katz & Mulawka (1968). Copper concentrations tended to be higher during the spring and summer suggesting a possible correlation with food uptake or temperature and the nature of resultant metabolism. Shiber (1979) notes seasonal variations in the copper content of prawns (Palaemon elegans); it was high in summer and spring and low in the fall. Levels in Crassostrea madrasensis from the Cochin region, India, were highest from December to May with low values confined to June to November (Sankaranarayanan et al., 1978). The authors believe freshwater river discharges to be responsible for the period of low copper burden. Betzer & Pilson (1974) found a 17-fold seasonal variation in the copper concentration of the blood of whelks which may reflect changes in copper availability or may be a physiological response related to the animal’s annual cycle of spawning activity. Regional differences in copper concentration have also been shown (Anderlini, 1974;

Interlinking of physical

1170

Wenzloff et al., 1979). Finally, consideration must be given to the fact that the coppercontaining protein, haemocyanin, is a respiratory pigment in crustaceans and molluscs (Bryan, 1968) and studies show that the blood as well as the hepatopancreas of crustaceans contain high levels of copper (e.g. Bryan, 1968; Wright, 1976). Nektonic Organisms (Table XI) Data on numerous fish species is summarized by Vinogradov (1953). Hall et al. (1978) provide a detailed survey of trace element levels of fish from 198 sites around the coasts of U.S.A. and the Department Fishery Data Centre, FAO (1978) provides a list of what results are available for fish. Copper levels in the present review range from below detectable limits to 51 µg·g dry wt−1 in whole fish to 148.1 µg·g dry wt−1 in the livers.

TABLE XI Levels of copper in selected marine nektonic organisms: units unless stated otherwise are mg/kg dry wt; m., mean; SD, standard deviation; ND, not detectable. *from a review paper, not from original

Species Fish (general)

Level

Comments

1–7 mg/kg

Netherlands

0.7–10

–

Reference Hueck, 1975

198 sites around coastal U.S.A. Hall et al., 1978 0.4–2.0 mg/kg

Whole

0.1–2.0 mg/kg

Muscle

1.0–110.0 mg/kg

Liver

Albacore

Goldberg, 1962 0.059% ash wt (CuO) Heart 0.009 ”

Spleen

0.008 ”

Gall bladder

0.098 ”

Liver

0.038 ”

Pyloric caeca

0.013 ”

Stomach

0.002 ”

Gill filaments

0.001 ”

Hyoid arch bone

0.0009 ”

Eyeballs

0.0013 ”

Integument

0.002 ”

Operculum bone

Oceanography and marine biology

0.0075 ”

Dorsal flesh

0.075 ”

Midline flesh

0.008 ”

Bone

697

Atlantic croaker 2.6

San Antonio Bay, Texas

Sims & Presley, 1976

Bay anchovy

2.8

San Antonio Bay, Texas

”

10.0

S.E.U.S.A. coast

”

Firth of Clyde, U.K.

Halcrow et al., 1973

Cod 2.3

Muscle

0.9

Liver

m. 0.47 mg/kg wet wt England and Wales, coastal

Species Cod

Dover sole

Level

Comments

m. 0.65 mg/kg wet wt

North Sea

m. 1.00 mg/kg wet wt

Distant waters (Iceland, Barents Sea, Norway)

0.6

Palos Verdes, CA Flesh—outfall (municipal wastewater)

0.5

Control

10.7

Gonads—outfall

12.7

Control

7.4

Liver—outfall

9.5

Control

Flounder

Firth of Clyde, U.K. 3.0

Muscle

0.8

Liver

Portmann, 1972

Reference Portmann, 1972

McDermott et al., 1976

Halcrow et al., 1973

Gilt-head bream

1.20 mg/kg wet Mediterranean wt

Ramelow et al., 1978

Gizzard shad

4.0

Sims & Presley, 1976

Grey mullet

1.70 mg/kg wet Mediterranean wt

Haddock

San Antonio Bay, Texas

Firth of Clyde, U.K. 2.2

Muscle

Ramelow et al., 1978 Halcrow et al., 1973

Interlinking of physical

1.1 Horse mackerel

Liver

0.99 mg/kg wet Mediterranean wt

Long rough dab

Menhaden

1172

Firth of Clyde, U.K.

Ramelow et al., 1978 Halcrow et al., 1973

2.3

Muscle

15.0

Liver

2.8

San Antonio Bay, Texas

Sims & Presley, 1976

Firth of Clyde, U.K.

Halcrow et al., 1973

Norway pout 2.6

Muscle

1.3

Liver

Plaice

Firth of Clyde, U.K. 2.3

Muscle

0.2

Liver

m. 0.85 mg/kg wet wt

England and Wales

m. 0.85 mg/kg wet wt

Muscle, coastal North Sea

m. 1.50 mg/kg wet wt

Distant water (Iceland, Barents, Norway)

Saithe

Firth of Clyde, U.K. 2.9

Muscle

0.3

Liver

”

Portmann, 1972

Halcrow et al., 1973

Sardine

2.18 mg/kg wet Mediterranean wt

Ramelow et al., 1978

Silverside

4.3

San Antonio Bay, Texas

Sims & Presley, 1976

Southern flounder

2.2

San Antonio Bay, Texas

Sims & Presley, 1976

3.0

S.E.U.S.A. coast

„

2.3

San Antonio Bay, Texas

”

8.4

S.E.U.S.A. coast

”

1.3

San Antonio Bay, Texas

”

Spot

Spotted sea trout

Oceanography and marine biology

9.5

699

S.E.U.S.A. coast

Species

Level

”

Comments

Reference

Striped mullet

0.68 mg/kg wet wt

Mediterranean

Ramelow et al., 1978

White bream

1.11 mg/kg wet wt

”

”

White sea trout

1.9

San Antonio Bay, Texas

Sims & Presley, 1976

Firth of Clyde, U.K.

Halcrow et al., 197

Whiting

Yellowfin tuna

Acanthopagarus australis

2.4

Muscle

0.7

Liver

0.040% ash wt (CuO)

Blood

0.0012 ”

Bone

0.024 ”

Pyloric caeca

0.048 ”

Eyeballs

0.014 ”

White flesh

0.050 ”

Dark flesh

0.004 ”

Gills

0.12 ”

Heart

0.12 ”

Intestines

0.09 ”

Intestinal contents

0.062 ”

Liver

0.016 ”

Integument

0.033 ”

Spleen

0.020 ”

Stomach

0.1–2.0 mg/kg

New South Wales, Australia

Goldberg, 1972

Bebbington et al., 1977

(m. 0.50) Acipenser transmoritanus

ND–0.97 mg/kg wet Lower Fraser River wt (m. 0.30) (estuarine), Canada

Northcote et al., 1975

Alosa sapidissima

0.22–0.23 mg/kg wet wt

Vinogradov, 1953*

Ammodytes lancea

m. 0.98, SD 0.33 mg/kg

From Belgian inshore fishery

De Clerck et al., 1979

A.lanceolatus

m. 0.92, SD 0.25 mg/kg

From Belgian inshore fishery

”

Interlinking of physical

1174

A.tobianus

m. 3.8

Cornwall, U.K.

Stevens & Brown, 1974

Anabas testudineus

0.162 mg/kg wet wt

Anarchicas lupus

2

Norwegian fjords

Lande, 1977

Anchoa mitchelli

10.0

Western North Atlantic

Windom et al., 1973

Anchoviella indica

1.04–1.84 mg/kg wet Strait of Singapore wt (m. 1.12)

Bok & Keong, 1976

1.55–2.16 mg/kg wet Strait of Malacca wt (m. 1.68)

”

Ancylopsetta quadrocellata

5.2

Stickney et al., 1975

Anguilla anguilla

0.17 µg/kg wet wt

Vinogradov, 1953*

Georgia estuaries

Vinogradov, 1953*

0.3–1.0 mg/kg wet wt

Lower Medway estuary, U.K. muscle

5.0–36.9 mg/kg wet wt

Liver Medway estuary, U.K. Muscle

Wharfe & Van Den Broek, 1977

Van Den Broek, 1979

0.25–6.99 mg/kg wet Small <30cm wt 0.4–1.25 mg/kg wet wt

Large >30 cm

0.84–4.99 mg/kg wet Liver (<30 cm) wt 1.35–19.90 mg/kg wet wt

Species

Level

Liver (>30 cm)

Comments Western North Atlantic

Reference

A.rostrata

0.8

Windom et al., 1973

Anistremus davidsoni

0.62 mg/kg wet Salton Sea, CA wt

Young & Mearns, 1978

Argentina silus

1

Norwegian fjord

Lande, 1977

Argyropelecus hemigymnus

0.949–1.807 mg/kg

Strait of Messina, Italy

Calapaj et al., 1978

Arripis georgianus

0.003 mg/kg wet wt

A.trutta

0.2–1.7 mg/kg

Vinogradov, 1953* New South Wales, Australia Bebbington et al., 1977

Oceanography and marine biology

701

(m. 0.87) 0.01 mg/kg wet wt

Vinogradov, 1953*

Atherinops affinis

0.20 mg/kg wet Newport Bay, CA wt

Young & Mearns, 1978

Bagre marinas

0.7

Western North Atlantic

Windom et al., 1973

Bairdiella chrysura

3.5

Western North Atlantic

”

2.3 ± 1.7

Georgia estuaries

Stickney et al., 1975

B.icistria

0.46 mg/kg wet Salton Sea, CA wt

Young & Mearns, 1978

Belone belone

m. 3.0–5.1

Cornwall, U.K., (2 locations)

Stevens & Brown, 1974

Ariake Sea

Enomoto & Toyodome, 1978

Boleophthalamus pectinirostris 2.17–3.24

Flesh

2.71–6.63

Internals

Carcharhinus falciformis

C.milberti

Western North Atlantic 2.1

Muscle

4.9

Liver

5.7

Kidney

Windom et al., 1973

8.4

Brain

<1.0

Gonads

6.3

Gills

4.6

Spleen

2.7

Western North Atlantic; liver

Windom et al., 1973

Western North Atlantic

”

C.obscurus 1.5

Muscle

1.3

Liver

8.4

Brain

1.6–2.4

Pup (total)

Carcharias sorrakowah

15.8–23.2

Goa Coast, India

Caranx georgianus

0.004 mg/kg wet wt

Catestomus macrocheilus ND–100 mg/kg Lower Fraser River wet wt (estuarine), Canada

Zingde et al., 1976 Vinogradov, 1953* Northcote et al., 1975

Interlinking of physical

1176

(m. 0.47) Centropristes striatus

<0.3

Western North Atlantic

Windom et al., 1973

Ceratoscopelus maderensis

1.070–1.644 mg/kg

Strait of Messina, Italy

Calapaj et al., 1978

C.warmingii

2.2

Western North Atlantic

Windom et al., 1973

Chauliodus sloani

1.120 mg/kg

Strait of Messina, Italy

Calapaj et al., 1978

Chimarera monostra

1

Norwegian fjord

Lande, 1977

Chrysophrys auratus

0.2–1.5 mg/kg

New South Wales, Australia Bebbington et al., 1977

(m. 0.59) C.guttulatus

0.01 mg/kg wet wt

Species Citharichthys sordidus

Vinogradov, 1953*

Level 0.19mg/kg wet wt

Comments

Reference

Palos Verdes, Calif.

Young & Mearns, 1978

S.California waste

Jan et al., 1977

0.20 mg/kg wet wt water outflow, JWPCP 0.09 mg/kg wet wt Orange county 0.17 mg/kg wet wt Island control C.spilopterus

1.2

Georgia estuaries

Clupea harengus

Stickney et al., 1975 Vinogradov, 1953*

1.42

Liver

0.23

Skin

0.23

Flesh

0.22

Gonads

4

Norwegian fjord

Lande, 1977

C.pilchardus

0.4 mg/kg wet wt

Vinogradov, 1953*

C.sardinea

1.82 mg/kg wet wt

”

S.sprattus

Northumberland coast, U.K. 14.2 ±4.1 mg/kg wet wt

Axial skeleton

8.88 ±5.0 mg/kg wet wt

Skin

4.61 ±4.07 mg/kg

Muscle

Wright, 1976

Oceanography and marine biology

703

wet wt Conger sp.

8.7

Western North Atlantic

Windom et al., 1973

Cottus asper

ND–0.74 mg/kg

Lower Fraser River (estuarine), Canada

Northcote et al., 1975

Northumberland coast, U.K.

Wright, 1976

(m. 0.46) Cyclopterus lumpus

1.13 mg/kg wet wt Skin 1.95 mg/kg wet wt Stomach wall 2.33 mg/kg wet wt Liver 1.3 mg/kg wet wt

Fat body

1.2 mg/kg wet wt

Kidney

0.76 mg/kg wet wt Muscle Cymatogaster aggregata

Puget Sound, U.S.A. 30.2

Dry ashing

5.0

Wet ashing

1.2 mg/kg wet wt

Wet ashing

Stober et al., 1977

Cynoscion atelodus

0.01 mg/kg wet wt

Vinogradov, 1953*

C.nebulosus

9.5

Western North Atlantic

Windom et al., 1973

C.regalis

1.9±0.8

Georgia estuaries

Stickney et al., 1975

C.xanthulus

0.30 mg/kg wet wt Salton sea, CA

Young & Mearns, 1978

Diplodus vulgaris

4.2

Mediterranean coast of Israel

Roth & Hornung, 1977

Dorosoma petenense 1.3 mg/kg wet wt

Salton Sea, CA

Young & Mearns, 1978

Elasmobranchii

1.5–3.2

N.Atlantic, Canada

Eustace, 1974*

Engraulis encrasicholus

0.5 mg/kg wet wt

Vinogradov, 1953*

1.8–3.1 mg/kg wet Adriatic; skin wt

Gilmartin & Revelante, 1975

0.6–0.9 mg/kg wet Gills wt 0.6–0.8 mg/kg wet Muscle wt 1.5–4.3 mg/kg wet Digestive tract wt

Interlinking of physical

1178

1.0–8.5 mg/kg wet Liver wt 0.96–1.52 mg/kg wet wt

Species

Total fish

Level

Comments

Reference

Engraulis encrasicholus

m., 1.18 SD 0.220 mg/kg

From Belgian inshore fishery

De Clerck et al., 1979

Epinephelus aeneus

54

Mediterranean coast of Israel

Roth & Hornung, 1977

E.ergastularius

0.01 mg/kg wet wt

E.guaza

3.3

Mediterranean coast of Israel

Roth & Hornung, 1977

E.guttatus

0.13–0.20 mg/kg

Gulf of Mexico, muscle

Taylor & Bright, 1973

Bahama Islands and Gulf of Mexico, muscle

”

E.striatus

Vinogradov, 1953*

0.11–1.15 mg/kg 9.91–110.20 mg/kg Liver Etmopterus spinax

6

Norwegian fjord

Lande, 1977

Etropus crossotus

0.9

Georgia estuaries, U.S.A.

Stickney et al., 1975

E.suratensis

14.9–17

Goa coast, India

Zingde et al., 1976

Strait of Malacca

Bok & Keong, 1976

1.28–4.39 mg/kg wet wt (m. 2.45)

Strait of Singapore

”

Euthynnus alletteratus

1.6

Western North Atlantic

Windom et al., 1973

Evynnis japonica

1.56

Japan

Ishii et al., 1978

Gadus aeglefinas

1.13

Eulamia melanoptera 1.31–1.62 mg/kg wet wt (m. 1.48)

1–3 G.merlangus

Norwegian fjord

0.33 2

G.morrhua

Vinogradov, 1953* Lande, 1977 Vinogradov, 1953* Norwegian fjord

0.041–0.38 mg/kg wet wt

Lande, 1977 Vinogradov, 1953*

1.5 mg/kg wet wt

Liver

”

3–4

Norwegian fjord

Lande, 1977

Oceanography and marine biology

705

Northumberland coast, U.K. 0.5 mg/kg wet wt

Gill

6.8 mg/kg wet wt

Axial skeleton

3.22 mg/kg wet wt

Skin

2.53 mg/kg wet wt

Stomach wall

2.9 mg/kg wet wt

Liver

1.0 mg/kg wet wt

Muscle

Wright, 1976

Galeorhinus australis 0.2–0.6 mg/kg wet wt

S.E.Australia

Glover, 1979

Galeus melastromus

Norwegian fjord

Lande, 1977

S.California waste water outflow, JWPCP

Jan et al., 1977

3

Genyonemus lineatus 0.21 mg/kg wet wt 0.17 mg/kg wet wt

Orange County

0.11 mg/kg wet wt

Coastal control

Girella tricuspidata

0.003 mg/kg wet wt

Glyptocephalus cynoglossus

1

Hemifuscus ternatanus

Vinogradov, 1953* Norwegian fjord

Lande, 1977

Ariake Sea

Enomoto & Toyodome, 1978

10.51–24.18

Flesh

21.36–239.68

Internals

Hexagrammos otakii 1.23

Japan

Holocentrus rufus

Texas outer Continental Shelf, Rezak et al., 1978

Species Holocentrus rufus

Level

Ishii et al., 1978

Comments

Reference

0.8

Muscle

28

Liver

3.2

Gills

Hygophum benoiti

1.508–2.220 mg/kg

Strait of Messina, Italy

Calapaj et al., 1978

H.hygomi

3.4

Western North Atlantic

Windom et al., 1973

Lactarius lactarius

2.6–5.4

Goa coast, India

Zingde et al., 1976

N.E.Atlantic

Stevens & Brown, 1974

Lamna nasus 18.1

liver

Rezak et al., 1978

Interlinking of physical

Lampanyctus pusillus

1180

6.0

Epigonal organ

2.7–23.0

Western North Atlantic

Lateolabrax japonicus 3.708 mg/kg wet wt 1.55 Lates calcarifer

Windom et al., 1973 Vinogradov, 1953*

Japan

0.02 mg/kg wet wt

Ishii et al., 1978 Vinogradov, 1953*

Leiognathus splendens 2.3–6.1

Goa coast, India

Zingde et al., 1976

Leiostomus xanthurus

1.8±0.9

Georgia estuaries

Stickney et al., 1975

8.4

Western North Atlantic

Windom et al., 1973

Lepidophanes indicus

13.0

”

”

Leptocottus armatus

0.23–0.62 mg/kg wet wt (m. 0.33)

Lower Fraser River (estuarine), Canada

Northcote et al., 1975

Limanda limanda

m. 0.44, SD 0.156 From Belgian inshore mg/kg fishery

L.ferruginea

New York Bight and Long Island Sound

De Clerck et al., 1979 Greig & Wenzloff, 1977

0.3–1.2 mg/kg wet Muscle wt 0.6–5.3 mg/kg wet Liver wt Lobianchia dofleini

23.0

Lopholatilus chamaeleonticeps

Western North Atlantic

Windom et al., 1973

New Jersey coast,

Mears & Eisler, 1977

80–320 mg/kg ash Males, liver 140–590 mg/kg ash Lutjanus blackfordi

0.038–0.16 mg/kg wet wt

L.campechanus

Makaira indica

Females, liver Vinogradov, 1953* Texas outer Continental Shelf,

0.8+0.3

Muscle

20+15

Liver

2.4–2.2

Gills

0.3–1.2 mg/kg wet N.E.Australia, muscle wt 0.5–22 mg/kg wet Liver wt

Rezak et al., 1978

MacKay et al., 1975a

Oceanography and marine biology

707

Macrurus rupestris

2

Norwegian fjord

Macullochella macquariensis

0.01 mg/kg wet wt

Vinogradov, 1953*

Malanogrammus aeglefinus

0.041 mg/kg wet wt

”

Maurolicus muelleri

1.793 mg/kg

Merlangius merlangus

Strait of Messina, Italy

Lande, 1977

Calapaj et al., 1978

Lower Medway estuary, Wharfe & Van Den U.K.; muscle Broek, 1977 0.2–1.0 mg/kg wet wt 1.3–3.5 mg/kg wet Liver wt 0.5–2.5 mg/kg wet Gut wall wt Medway estuary, U.K. 0.3–0.75 mg/kg wet wt

Muscle (small <14.9 cm)

ND–0.46 mg/kg wet wt

Muscle (large >14.9 cm)

Species

Level

Van Den Broek, 1979

Comments

Merlangius merlangus ND–2.49 mg/kg wet Liver wt

Reference Van Den Brock, 1979

0.46–1.30 mg/kg wet Gut wall wt Merluccius bilinearis

0.1 mg/kg wet wt

M.merluccius

2.9–5.2

M.vulgaris

0.6 mg/kg wet wt

Micropogon undulatus 2.3

Vinogradov, 1953* Mediterranean coast of Israel

Vinogradov, 1953* Georgia estuaries

0.25 mg/kg wet wt Microstonus pacificus 1.1–9.2 mg/kg wet wt Mola sp.

0.14 mg/kg wet wt

Roth & Hornung, 1977

Stickney et al., 1975 Vinogradov, 1953*

Los Angeles, liver

De Goeij et al., 1973 Vinogradov,

Interlinking of physical

1182

1953* Monacanthus chinensis

0.004 mg/kg wet wt

Morone saxatilis

2.5

Western North Atlantic

Windom et al., 1973

0.27 mg/kg wet wt

Newport Bay, CA

Young & Mearns, 1978

1.9

Western North Atlantic

Windom et al., 1973

Mugil cephalus

”

0.003–0.82 mg/kg wet wt

Vinogradov, 1953*

0.24–0.55 mg/kg wet Newport Bay, CA wt

Young & Mearns, 1978

0.2–2.8 mg/kg

New South Wales, Australia

Bebbington et al., 1977

M.parsia

28.6–32.5

Goa coast, India

Zingde et al., 1976

Mulus barbatus

4.2–6.4

Mediterranean coast of Israel

Roth & Hornung, 1977

M.surmulletus

m. 0.68; SD 0.300 mg/kg

From Belgian inshore fishery

De Clerck et al., 1979

Mustelus antarcticus

0.004 mg/kg wet wt

M.canis

Vinogradov, 1953*

0.2–0.4 mg/kg wet wt

S.E.Australia

Glover, 1979

0.7–1.0 mg/kg wet wt

New York Bight and Long Island Sound, muscle

Greig & Wenzloff, 1977

0.6–1.5mg/kg wet wt Liver Mycteroperca phenox

0.10–0.36 mg/kg

Gulf of Mexico and Bahamas, Taylor & Bright, muscle 1973

M.tigris

0.11–1.17 mg/kg

Gulf of Mexico and Bahamas, ” muscle

34.60–65.53 mg/kg

Liver

0.13–0.33 mg/kg

Gulf of Mexico and Bahamas, Taylor & Bright, muscle 1973

M.venosa

Myctophum punctatum 1.604–3.334 mg/kg Mylocheilus caurinus

Strait of Messina, Italy

ND–1.53 mg/kg wet Lower Fraser River wt (m. 0.66) (estuarine), Canada

Calapaj et al., 1978 Northcote et al., 1975

Oceanography and marine biology

Myoxocephalus scorpius

709

Strathcona Sound, 1.6–9.9 (m. 4.1)

NWT, Canada; muscle

1.8–26 (m. 7.6)

Liver

Neoplatycephalus macrodon

0.003 mg/kg wet wt

Notoscopelus caudispinous

3.2

Oncorhynckus nerka

0.69–0.80 mg/kg wet Lower Fraser River wt (m. 0.73) (estuarine), Canada

Species

Bohn & Fallis, 1978

Vinogradov, 1953* Western North Atlantic

Level

Oncorhynchus nerka

Comments During migration to Chilko River, B.C.

Windom et al., 1973 Northcote et al., 1975

Reference Fletcher & King, 1978b

Lumni Island, sea water 8.07±1.02mg/100g body wt

Ovary

19.7±2.52 µg/100g body wt

Testes

7.39±1.25 mg/100g body wt

Liver (female)

14.1±1.80 mg/100g body wt

Liver (male) Chilko Lake, fresh water

O.tschawytscha

7.12±1.22 mg/100g body wt

Ovary

12.2±5.22 µg/100g body wt

Testes

4.22±0.73 mg/100g body wt

Liver (female)

10.2±4.35mg/100g body wt

Liver (male)

0.40–0.70 mg/kg wet wt (m. 0.55)

Lower Fraser River (estuarine), Canada

0.40 mg/kg wet wt Ophichthus gomesi

1.5

Northcote et al., 1975 Vinogradov, 1953*

Western North Atlantic

Windom et al., 1973

Interlinking of physical

1184

O.ocellatus

2.7

”

”

Opsanus tau

1.8±0.8

Georgia estuaries

Stickney et al., 1975

Orthopristus chrysopterus

ND

Otolithus ruber

4.9–12.6

Goa coast, India

Zingde et al., 1976

Pampus argenteus

3.5–12

”

”

S.California waste water outflow,

Jan et al., 1977

Vinogradov, 1953*

Paralabrax clathratus

P.maculatofasciatus

0.19mg/kg wet wt

Orange County

0.13 mg/kg wet wt

Island control

0.26 mg/kg wet wt

Newport Bay, CA

Young & Mearns, 1978

S.California waste water outfall,

Jan et al., 1977

Paralichthys californicus 0.1 3 mg/kg wet wt

JWPCP

<0.2 mg/kg wet wt

Coastal control

P.lethostigma

3.0

Western North Atlantic

Windom et al., 1973

P.olivaceus

0.60

Japan

Ishii et al., 1978

Parophrys vetulus

5.3

Puget Sound, U.S.A., dry Stober et al., ashing 1972

Pelates sexlineatus

0.004 mg/kg wet wt

Pellona ditchela

3.6–8.1

Vinogradov, 1953*

Platichthys flesus

Species Platichthys flesus

Goa coast, India

Zingde et al., 1976

Medway estuary, U.K.

Van Den Broek, 1979

ND–1.33 mg/kg wet wt

Muscle

ND–23.01 mg/kg wet wt

Liver

1.00–4.78 mg/kg wet wt

Gut Wall

Level

Comments

Reference

ND–0.12 mg/kg wet wt

Ovary

Van Den Broek, 1979

m. 0.45, SD

From Belgian inshore

De Clerck et al., 1979

Oceanography and marine biology

Platycephalus fuscus

0.091 mg/kg

fishery

0.1–1.3 mg/kg

New South Wales, Australia

711

Bebbington et al., 1977

(m. 0.47) Plectroplites ambiguus

0.004 mg/kg wet wt

Pleuronectes limanda

Vinogradov, 1953* Norwegian fjord

4

Flesh

10

Gill

P.platessa

Northumberland coast, U.K. 2.14 mg/kg wet wt

Stenner & Nickless, 1974a

Wright, 1976

Axial skeleton

3.8 mg/kg wet wt Skin 4.2 mg/kg wet wt Stomach wall 1.89 mg/kg wet wt

Liver

0.5 mg/kg wet wt Muscle Lower Medway estuary, Wharfe & Van Den U.K. Broek, 1977 0.5–1.7 mg/kg wet wt

Muscle

1.5–4.2 mg/kg wet wt

Liver

1.8–3.2 mg/kg wet wt

Gut Wall Medway estuary, U.K.

ND–1.40 mg/kg wet wt

Muscle

ND–3.74 mg/kg wet wt

Liver

ND–1.89 mg/kg wet wt

Gut wall

Van Den Broek, 1979

Poecilia latipinna

0.3 mg/kg wet wt Salton Sea, CA

Young & Mearns, 1978

Polyprion americanus

0.4 mg/kg wet wt

Vinogradov, 1953*

Pomatomus pedica

0.01 mg/kg wet wt

”

Interlinking of physical

P.saltatrix

1186

0.44–0.58 mg/kg Atlantic, white muscle wet wt

Cross et al., 1973, 1975

0.2–1.4 mg/kg

New South Wales, Australia

Bebbington et al., 1977

New Jersey coast

Mears & Eisler, 1977

(m. 0.67)

160–820 mg/kg ash wt

Males, liver

150–1600 mg/kg Females, liver ash wt Pomatoschistus minutus

0.5–1.5 mg/kg wet wt

Potamalosa novaehollandiae

0.01 mg/kg wet wt

Prionace glauca

Pseudomonacanthus ayraudi

Lower Medway estuary, Wharfe & Van Den U.K. Broek, 1977 Vinogradov, 1953* Cornwall, U.K.

m. 5.7

Liver

m. 4.4

Muscle

m. 5.6

Gonad and epigonal organ

0.004 mg/kg wet wt

Species

Ptychocheilus oregonensis

Vinogradov, 1953*

Level

Pseudopleuronectes americanus

Stevens & Brown, 1974

Comments New York Bight and Long Island Sound

Reference Greig & Wenzloff, 1977

0.5–1.1 mg/kg wet wt

Muscle

2.7–13.8 mg/kg wet wt

Liver

ND–1.06 mg/kg

Lower Fraser River (estuarine), Canada

Northcote et al., 1975

Western North Atlantic

Windom et al., 1973

(m. 0.63) Raja eglanteria

Rastrelliger kanagurta

3.2

Muscle

44

Liver

4.4

Yolk sac

15.8

Goa coast, India

Zingde et al., 1976

Oceanography and marine biology

Rastrelliger sp.

713

1.15–1.98 mg/kg wet Strait of Singapore wt (m. 1.33)

Bok & Keong, 1976

2.16–3.14 mg/kg wet Strait of Malacca wt (m. 2.60)

”

Regificola grandis

0.01 mg/kg wet wt

Vinogradov, 1953*

Reporhamphus australis

0.002 mg/kg wet wt

”

Rhinobatis lentiginous

Western North Atlantic 2.2

Muscle

6.6

Liver

6.2

Stomach

2.7 Rhinoptera bonusus

Yolk sac Western North Atlantic

2.3

Muscle

13

Liver

10

Brain

7.0

Stomach

5.2

Spiral

3.6

Spleen

3.4

Uterus

Rhomboplites aurorubenes

Windom et al., 1973

Texas outer Continental Shelf, muscle

”

Rezak et al., 1978

1.0±0.2 31±32

Liver

3.4±2.8

Gills

Roboralga jacksoniensis

0.01 mg/kg wet wt

Vinogradov, 1953*

Roughleyia australis

0.003 mg/kg wet wt

Vinogradov, 1953*

Salmo clortii

0.049–0.90 mg/kg wet wt (m. 0.66)

S.gairdneri

0.22–1.02 mg/kg wet Lower Fraser River wt (estuarine), Canada

”

S.malma

0.30–0.82 mg/kg wet Lower Fraser River wt (m. 0.56) (estuarine), Canada

”

Sardina pilchardus

6.6

Cornwall, U.K.

Stevens & Brown, 1974

N.Adriatic

Gilmartin & Revelante, 1975

Lower Fraser River (estuarine), Canada

Northcote et al., 1975

Interlinking of physical

1188

0.9–1.8 mg/kg wet wt

Skin

0.6–1.1 mg/kg wet wt

Gills

0.9–1.2 mg/kg wet wt

Muscle

0.8–1.7 mg/kg wet wt

Digestive tract

2.1–3.7 mg/kg wet wt

Liver

0.94–1.09 mg/kg wet Total fish wt

Species

Level

Comments

Reference

Sardinella aurita

2.8–6.0

Mediterranean coast of Israel

Roth & Hornung, 1977

S.fimbriata

5.9–12.4

Goa coast, India

Zingde et al., 1976

Saradinia caerlea

0.166 mg/kg wet wt

Saurida undosquamis 0.7–6.4 Sciaena amarctica

Vinogradov, 1953* Mediterranean coast of Israel

0.01 mg/kg wet wt 0.1–2.4 mg/kg

Roth & Hornung, 1977 Vinogradov, 1953*

New South Wales, Australia

Bebbington et al., 1977

(m. 0.64) S.coitor

0.053 mg/kg wet wt

Vinogradov, 1953*

S.maculata

0.029 mg/kg wet wt

”

Scillium caniculata

0.43

”

Scomber australasicus

0.003 mg/kg wet wt

”

S.scombrus

0.115–0.86 mg/kg wet wt

”

5.5

Cornwall, U.K.

Stevens & Brown, 1974

Scomberomorus maculatus

2.3

Western North Atlantic Windom et al., 1973

Scomberomorus sp.

0.79–2.23 mg/kg wet wt (m. 1.34)

Strait of Singapore

Bok & Keong, 1976

1.74–2.26 mg/kg

Strait of Malacca

”

Oceanography and marine biology

715

wet wt (m. 1.95) Scophthalmus aquosus

2.0

Scorpaena guttata

Sebastes nivosus

Georgia estuaries

Stickney et al., 1975

S.California waste water outflow

Jan et al., 1977

0.15 mg/kg wet wt

JWPCP

0.10 mg/kg wet wt

Orange County

0.15 mg/kg wet wt

Coastal control

0.15 mg/kg wet wt

Palos Verdes, CA

Young & Mearns, 1978

0.51

Japan

Ishii et al., 1978

S.California waste water outflow,

Jan et al., 1977

S.paucispinis 0.15 mg/kg wet wt

JWPCP

0.13 mg/kg wet wt

Island control

S.auriculatus

Santa Barbara Channel

McDermott-Ehrlich & Alexander, 1976

5.0–9.4

Liver

3.0–4.2

Kidney

S.thompsoni

0.87

Japan

Ishii et al., 1978

Selar sp.

1.54–1.59 mg/kg wet wt (m. 1.56)

Strait of Singapore

Bok & Keong, 1976

1.77–1.91 mg/kg wet wt (m. 1.80)

Strait of Malacca

”

Strait of Singapore

”

0.77–1.20 mg/kg wet wt (m. 1.0)

Strait of Malcca

”

0.2–1.7 mg/kg

New South Wales, Australia

Bebbington et al., 1977

0.85–2.42 mg/kg wet wt (m. 1.18)

Strait of Singapore

Bok & Keong, 1976

1.22–1.31 mg/kg wet wt (m. 1.26)

Strait of Malacca

”

S.rivulatus

3.1

Mediterranean coast of Israel

Roth & Hornung, 1977

Solea solea

1.4

Mediterranean coast of

”

Selaroides leptolepis 0.89–2.20 mg/kg wet wt (m. 1.32)

Seriolia grandis

(m. 0.59) Siganus oramin

Interlinking of physical

1190

Israel

Species

Level

Sphyraena novaehollandiae

0.002 mg/kg wet wt

S.sphyraena

4.8–23.5

Sphyrna lewini

Reference Vinogradov, 1953*

Mediterranean coast of Roth & Hornung, 1977 Israel Western North Atlantic Windom et al., 1973

2.0

Muscle

6.2

Liver

10

Stomach

5.7

Intestine

S.tiburo

Sprattus sprattus

Comments

Western North Atlantic Windom et al., 1973 3.0

Muscle

3.6

Liver

9.2

Stomach

2.4

Spleen

2.4

Ovary

0.5–2.0 mg/kg wet wt Lower Medway estuary, U.K.

Squalus acanthius

Wharfe & Van Den Broek, 1977

Western North Atlantic Windom et al., 1973 2.3

Muscle

4.5

Liver

4.8

Stomach

16

Spleen

0.9

Yolk sac

3.0

Embryo

S.mitsukurii

Choishi, Japan 0.13–0.56 µg/g wet wt (m. 0.30)

Whole, male

0.12–0.70 µg/g wet wt (m. 0.26)

Whole, female

0.23–0.65 µg/g wet wt (m. 0.45)

Embryo (muscle)

0.45–0.55 µg/g wet

Age 0

Taguchi et al., 1979

Oceanography and marine biology

717

wt (m. 0.48) Stellifer lanceolatus

1.8±0.8

Georgia estuaries

Stickney et al., 1975

Symphurus plagiusa

1.6±1.0

”

”

Synaptura nigra

0.003 mg/kg wet wt

Vinogradov, 1953*

Tarpon atlanticus

0.24 mg/kg wet wt

” New Jersey coast

Mears & Eisler, 1977

430–1 350 mg/kg ash Male, liver wt

Thunnus albacares

280–3700 mg/kg ash wt

Female, liver

0.2–0.7 mg/kg

New South Wales, Australia

Bebbington et al., 1977

Cadiz Coast, Spain

Establier, 1970b

T.thynnus m. 148.1

Liver

m. 16.6

Heart

m. 7.0

Intestine

m. 4.9

Gills

m. 23.8

Kidney

m. 60.8

Pancreas

m. 7.1

Stomach

m. 4.1

Spleen

Trachichthodes affinis 0.002 mg/kg wet wt

Species Trachurus trachurus

Level

Vinogradov, 1953*

Comments

Reference

m. 0.56 SD 0.133 mg/kg

From Belgian inshore fishery

De Clerck et al., 1979

4.0

Cornwall, U.K.

Stevens & Brown, 1974

Trigla lucerna

m. 0.70 SD 0.1 34 mg/kg

From Belgian inshore fishery

De Clerck et al., 1979

Trisopterus luscus

m. 0.57 SD 0.130 km/kg

From Belgian inshore fishery

”

Umbrina roncadar 0.26 mg/kg wet wt

Newport Bay, CA

Young & Mearns, 1978

Upeneus moluccensis

Mediterranean coast of Israel

Roth & Hornung, 1977

4.5–8.3

Interlinking of physical

Urophycis chuss

1192

New York Bight and Long Island Sound 0.5–0.7 mg/kg wet wt

Muscle

2.7–6.0 mg/kg wet wt

Liver

Vinciguerria attenuata

1.295 mg/kg

Strait of Messina, Italy

Zeus australis

0.002 mg/kg wet wt

Zoarces viviparus

Greig & Wenzloff, 1977

Calapaj et al., 1978 Vinogradov, 1953*

Northumberland, U.K.

Wright, 1976

5.55±5.0 mg/kg wet Axial skeleton wt 5.6±3.5 mg/kg wet wt

Skin

2.3 mg/kg wet wt

Stomach wall

5.18 mg/kg wet wt

Liver

0.77±0.44 mg/kg wet wt

Muscle

Eustace (1974) found no correlation between copper concentrations and length of a number of deep water fish. Cross et al. (1973) felt copper levels remained constant or decreased with the size of bluefish (Pomatomus saltatrix). Mears & Eisler (1977) found that copper levels decrease as the size of Tautog onitus increases. Bohn & Fallis (1978) showed that copper concentrations in the liver and muscle tissue of Myxocephalus scorpius increased as the body weight increased. This is supported by Mears & Eisler (1977) who found that larger Lopholatilus chamaeleonticeps contained more copper in liver tissues than did smaller fish of the same species. Other authors have found, however, that the copper concentration in muscle tissues decreases as size increases (Taguchi, Yasuda, Toda & Shimizu, 1979; Van den Broek, 1979). In light of their results, Mears & Eisler (1977) suggest that comparisons for environmental studies should only be made using fish of the same age. Goldberg (1962) showed that metals which form strong organic complexes tended to be concentrated in internal organs. Establier (1970a), Establier & Gutierrez (1970) and Vinogradov (1953), for example, showed copper to be concentrated in the liver, pancreas, kidney, and spleen of tunas. Copper concentrations will also vary depending upon the environmental conditions which exist and the metal levels encountered in food organisms as well as the physiological state of the organism. Establier (1970a) and Establier & Gutierrez (1970) showed a distinct seasonal difference in the copper content of tuna. De Clerck, Vanderstappen, Vyncke & Van Hoeyweghen (1979) found that samples of fish caught in Belgian coastal waters were similar to those caught in open sea in terms of metal content. Elimination of metal tends to be greater at higher temperatures but salinity has no effect (Nakahara, Koyanagi & Saiki, 1972).

Oceanography and marine biology

719

Several authors have noted changes in copper concentrations related to the life histories of necktonic organisms. Taguchi et al. (1979) found copper concentrations were higher in embryonic stages and immediately after the birth in Squalus mitsukurii than later stages. Copper concentrations in the blood plasma of rockeye salmon (Oncorhynchus nerka) decreased during the non-feeding migration to spawning grounds (Fletcher & King, 1978b). MARINE MAMMALS AND BIRDS These organisms can be said to exhibit basically the same tendency as invertebrates and members of the nekton, i.e. to concentrate heavy metals in digestive, excretory and, to a degree, reproductive tissues. Metabolic regulation of metal uptake does occur (e.g. Anderlini et al., 1972). TRANSFER OF COPPER THROUGH THE FOOD CHAIN One of the major concerns about copper and other metals in the marine environment is the possibility of concentration in the various levels of the food chain. The result would be that higher levels of the food chain would have higher concentrations of metals. This possibility formed one of the reasons for the CEPEX studies of pollutants in marine food chains (Menzel, 1977). The results (e.g. Beers et al., 1977; Koeller & Parsons, 1977) suggest that the primary effect of copper on marine food chains is not the concentration at higher trophic levels, but rather to change the species composition of the lower portion of the food chain. This in turn alters the spectrum of food items available. Another series of studies of a benthic food chain included algae, the bivalve Tellina tenuis, and Pleuronectes platessa (plaice) which feeds on the siphons of Tellina. Copper did not accumulate in the plaice muscle but did in the viscera and in the bivalve (Saward, Stirling & Topping, 1975; Topping, 1977). The copper also reduced the standing stock of algae and impaired siphon regeneration in Tellina. Pollutant flow through food webs was examined for the southern California Bight by Young & Mearns (1978) who state that “public apprehension regarding the accumulation of pollutants in seafood is based largely on the assumption that food chain magnification of organic and inorganic contaminants, which has been demonstrated in certain terrestrial and freshwater systems, also occurs in marine ecosystems…the evidence obtained to date indicates that there is measurable structure to the coastal marine ecosystems of the Southern California Bight. Despite this structure, concentrations of most trace metals of present concern decrease with increase in presumed trophic level.” The exception to this trend was mercury. Metal concentrations in a variety of different organisms indicate considerable interand intraspecific variability. Trace metal levels in beach dipterans and beach amphipods were found to be consistently higher in the dipterans (Bender, 1975). Andreev & Markov (1971) found accumulations of copper in intestinal parasites up to 30–40 times that in the intestine of the host. Hardisty, Kartar & Sainsbury (1974), working with zinc, lead, and cadmium found that, with lead and cadmium, concentration in fish was a function of the proportion of crustaceans in the diet while there was no relationship between diet and zinc content. Herbert, Wiener & Field (1978) examined the levels of copper in the liver,

Interlinking of physical

1194

kidney, and plasma of breeds of sheep fed copper-enriched meal containing some dried seaweed. They found a difference in copper accumulation in the liver between breeds and also found that increasing the percentage of the diet made up of seaweed (Ascophyllum nodosum) decreased the amount of copper accumulated. In examining the flow of copper through a terrestrial food chain including an isopod, Wieser (1961), Dallinger (1977), Dallinger & Wieser (1977), and Weiser, Dallinger & Busch (1977) noted that at “low” concentrations of copper in the food (20 µg·g−1), more copper was lost through faeces than ingested, while at higher levels metal assimilation increased to between 80–90% of the metal ingested. These authors also found that the isopods selected litter providing the appropriate copper level and that copper concentration occurred in the hepatopancreas. Cross et al. (1975) noted that metal assimilation efficiencies of juvenile fish in a coastal plain estuary were highly variable and that the efficiency changed with the trace metal composition of the food which varied with the distribution of the fish within the estuary. Aubert et al. (1975) found that the chemical species of the metals studied produced different concentration factors, that complexed copper caused lower accumulation in Nereis diversicolor, and that copper accumulation in Carcinus maenas was, at least in part, due to adsorption to the cuticle. The production of organic material requires the uptake of nutrients and trace metals. Aquatic plants accumulate trace elements from the water and, in some cases, from the bottom sediments (Varenko & Chuiko 1971). Animals incorporate trace elements directly from the water, indirectly from the food or from both. The movement of metals through the food chain, from the plant through the primary herbivore, secondary herbivore, carnivore, and detritus-feeder, is complex. Rozhanskaya (1969) indicates a general decline of copper in the higher levels of the food chain although organisms feeding on detrital material or living on or in bottom sediments are frequently exceptions. A second generality, which tends to simplify the understanding of the transfer of the metals, is the statement in Phelps et al. (1969) that “any element present in the body of an organism, whether it is incorporated into its biological system, or merely occurs as an ‘adventitious’ form passing through the gut, becomes significant for the next trophic level, and for the biogeochemical cycling of the element itself”. From this generalization it is possible to see how seasonal cycles can impose changes on the partitioning of various copper species in components of food webs. The transfer of metals through the food chain appears to be at least partially controlled by biological requirements for the metals. Thus, in a study of the transfer of 65Zn and 51Cr through an estuarine food chain (Baptist & Lewis, 1969), the transfer resulted in higher concentrations of 65Zn, a required metal, than 51Cr a metal required very little or not at all. The authors discuss the uptake of ionic versus particulate forms of metals as a possible explanation of the differences. Differential rates of transfer from food and dissolved metals were observed for post-larval Micropogon undulatus and Fundulus heteroclitus, higher rates of uptake occurring from the food than from the water. In discussing the transfer of metals through food chains it is necessary to consider environmental factors. Transfer of metals from unnatural sources is dependent upon the nature of the metal source (Baptist & Lewis, 1969). In addition, the effect of continuous input will not be the same as that of a single slug, accumulation will be greater with constant exposure (Baptist & Lewis 1969). Migration by the organism into and out of

Oceanography and marine biology

721

contaminated areas will also affect accumulation. Availability of the metal, whether from a natural or unnatural source, will be affected by particulates and organics (e.g. Siegal, 1966; Prakash & Rashid, 1968; Bender et al., 1970; Steemann Nielsen & WiumAnderson, 1970; Lewis et al., 1972, 1973). Other factors, discussed on pp. 507–522, will also affect the accumulation of metals. The concentration of metal in an organism is, in part, dependent upon the amount of food ingested as well as the level of the metal in the food. Because potential food organisms may accumulate metals differently during their life cycle, the age of the food organism may affect the concentration of metals. The concentration is also dependent upon the species and past history of food organism because different organisms tend to have different levels of metals (e.g. Ireland, 1973). Thus, although Ireland found that the zinc content of organisms eating seaweed and algae decreased as one proceeded away from a pollution source, the values in filter-feeding benthic organisms (Balanus balanoides and Mytilus edulis) were more variable. Likewise, the effect of a metal will vary with different organisms which makes the prediction of effect, at various levels in the food chain, difficult. Ruthven & Cairns (1973) for example, found differential effects of “toxic” levels of copper when presented to several species of organisms. Extrapolating from this, the overall effect of a potentially toxic level of copper may be to change the structure of marine communities. Removal or reduction in numbers of one or more components of a food chain may, however, cause a change in the other components due to food preferences (Koeller & Parsons, 1977). Young & Mearns (1978) have calculated transfer coefficients between trophic levels in ecosystems of the southern California Bight showing a decreased copper level with increased trophic level. This finding is echoed by Rozhanskaya (1969). Waldichuk (1974) calculates copper bioaccumulation factors of phytoplankton (38), zooplankton (437), macro-invertebrates (24 000–35 000), and fish (50–250). Eisler (1979a) provides a general review of copper accumulation in coastal marine biota. Many authors provide evidence of food chain concentration of metals (e.g. Patrick & Loutit, 1976; Robb, 1976; Harms et al., 1978; Neff et al., 1978; Bohn, 1979; Martin, 1979; Montgomery & Price, 1979; Vermeer & Peakall, 1979; Kamimura, 1980). Anderlini et al. (1972) found no significant geographical or species variation in four petrel populations of Antarctic and North American seas and suggest this to be a result of metabolic regulation of metal uptake. The interconnection between seasonal changes in metal concentrations, changes in concentration with ontogeny, and transfer of metals through the food chain can be shown by the study of Williams & Murdoch (1969) on Spartina alterniflora. This plant accumulates high levels of zinc, manganese, and iron, and forms an important source of detritus in some estuarine systems. Levels of these metals varied in concentration during seasonally related growth periods. Levels were lowest in mature plants and highest in the shoots. After the death of the plant, however, the metal levels increased markedly. This increase was suggested to be due to materials deposited by microflora or by some adsorption process (see Williams, Howell & Straub, 1960). Spartina is consumed mostly as detritus and the effect of the increase in metal concentrations after death would be expressed in the detritus-feeding organisms first. Introduction of Spartina detritus could thus provide a seasonal source of metals to detritus-feeders. Faecal material forms an important source of food for many animals (Harvey, 1963)

Interlinking of physical

1196

and trace metals taken up in the food have been shown to be concentrated in the faeces. Boothe & Knauer (1972) found that Macrocystis pyrifera contained an average of 6.9 µg Cu·g ash−1 while the faeces of the kelp crab Pugettia producta, which feeds on the alga, contained an average of 30 µg Cu·g ash−1. They estimated the concentration factor of copper (faeces: algae) to be 4.3 suggesting that coprophagic organisms could be faced with a relatively high level of copper (and other trace metals) in their food. Breakdown of the faecal material to detritus would, as in Spartina, provide a metal-rich food source to detritus-feeders. Benthic organisms are exposed not only to dissolved metals in the water, but also to metals adsorbed to the substratum or in interstitial water. A difference in the concentration of metal may be due to the effect of the substratum as well as the concentration of metal in the water and in the food. Thus, the levels of copper in organisms living in the delta clayey silt of the Añasco River of Puerto Rico were found to be significantly higher than those in organisms living in either relict sand and gravelly sand or basin clayey silt (Phelps et al., 1969). In this study, no significant difference was demonstrated between the accumulation of copper and the mode of feeding (‘non-select’ and ‘select’ deposit-feeders, filter-feeders, and carnivores). A significant difference was found, however, with zinc, iron, scandium, and samarium. There thus appears to be a response by the organism to different trace metals in the food and an apparent background level of metal established by environmental conditions.

LEVELS OF COPPER FOUND TO BE TOXIC TO MARINE ORGANISMS Copper toxicity results from the same mechanisms and properties which make the metal an essential constituent of many metalloproteins, its ability to enter into strong complexes with organic ligands. Copper may thus react with proteins to denature them. As a result, toxicity may be due to a reduction in enzyme activity or in the destruction or distortion of protein structure. The toxic effects of copper are often referred to as “chronic” or “acute”. Abnormalities in development, productivity, respiratory, feeding and growth rates, and burrowing or other behaviour which are the result of the detrimental activity of copper are said to be chronic. The reduction in the percentage of organisms hatching or surviving is an acute effect. Table XII is a compilation of most of the toxic levels of copper encountered in the literature. Review references pertaining to this topic include Davies (1978), an excellent review of theory and literature on pollution studies with marine plankton, and Eisler (1979b) and Hodson et al. (1979), both with reviews of the effects of heavy metals on marine biota, although the latter reference tends more towards freshwater conditions. Wittmann (1979) provides a general review of toxic metals including discussions of the reasons for the toxicity. An extensive annotated list of toxic concentrations of copper is given by Black et al. (1975). The primary sources of confusion in evaluating the toxicity values are the variety of toxicity indices, the variability produced by different experimental conditions, and the often indiscriminate use of toxicity terms. Definitions for some of the commonly encountered toxicity indices are given below where L=lethal (or level), M or m=mean or

Oceanography and marine biology

723

median, D=dose, C =concentration, and T=toxicity (or time). (1) LC50, MLD, MLC, TLm, TL50—The concentration resulting in a mortality of 50% of the organisms within a given length of time (most commonly 96 h LC50). LC0 and LC100 are concentrations resulting in 0 and 100% mortalities, respectively. (2) LT50, LTm, MLT, TLm*, TL50*—The length of time, at a given concentration, necessary to cause 50% mortality of the organisms. One will often encounter LT0 and LT100 indices indicating duration of 0 mortality and the length of time to produce 100% mortality respectively. (*Normally used with the concept given in item 1 although sometimes with the concept given in item 2, e.g., Heslinga, 1976, and Martin et al., 1977). Other, less commonly encountered indices include the following. (3) “Incipient Lethal Level” (ILL), that level beyond which the organism can no longer live for an indefinite period of time (Fry, 1947). (4) “Minimum Lethal Dose” (MLD), lowest concentration to which lethal effects can be attributed (Segar et al., 1971). Note that this differs from the MLD described above (1) where “M” stands for mean or median. (5) Threshold toxicity (=toxicity threshold)—the concentration required to elicit a toxic response (e.g., McLusky & Phillips, 1975; Davenport & Manley, 1978). (6) No effect level—the highest concentration which does not result in toxicity (e.g., Hueck, 1975.) (7) EC50—“Effective concentration” to yield 50% activity (i.e., enzyme activity or phytoplankton growth or motility, see Rosko & Rachlin, 1975 and Brown, 1976). (8) I50—the concentration at which 50% inhibition of photosynthesis occurs (e.g., Overnell, 1976). With the exception of EC50 and I50, the normal condition applied to these indices is the death of the bioassay organism. Many authors have applied other criteria to toxicity studies. Anderson & Morel (1978) note 100% non-motility of the dinoflagellate Gonyaulax tamarensis at a calculated cupric ion activity of 10−9.7 M (50% at 10−10.4M). A number of references are provided in Table XII which relate copper toxicity to the growth or photosynthetic capabilities of phytoplankton. In tests using the marine diatom Cylindrotheca closterium Lehman & Vasconcelos (1979) and Lehman (1979) use the photosynthetic rate as an index of copper toxicity noting that changes in the photosynthetic rate parallel changes in the respiratory rate. Lehman & Vasconcelos (1979) also found that metal concentrations of 100 µg·1−1 caused clumping of the cells, a change in shape from the normal fusiform to oval, morphological changes such as increased vacuolar volume, and considerable changes in the distribution of lipids in whole cells and chloroplasts (both polar and non-polar classes). The inhibition of photosynthesis by copper in populations of micro- and nannoplankton was found to be much more than other metals, however, levels of up to 8 and 4 µg Cu·1−1, respectively, were found to enhance productivity in these populations (Rajendran et al., 1978). Alayse-Danet et al. (1979), in studies of Artemia salina exposed to copper, noted that disturbances in the activity of the enzyme amylase occur more rapidly than decreases ingrowth rate while disturbances of trypsin activity occurred

Interlinking of physical

1198

simultaneously with growth rate decreases. EC50 values of 78×10−6M were found for the activity of the enzyme allantoinase in the polychaete Eudistylia vancouveri by Brown (1976) and of 0.130–0.190 mg·1−1 Cu for the growth rate of the diatom Nitzschia closterium over 96 h (Mandelli, 1969). Mueller’s (1979) study of some species inhabiting the intertidal flats of the Outer Elbe Estuary showed low copper concentrations (5 µg·1−1) affected oxygen consumption, caused changes in behaviour, and resulted in morphological modifications (mostly obstruction of the respiratory epithelia by mucus). Working with estuarine bacterial populations of the Humber Estuary Goulder et al. (1979b) used Vmax for the rate of glucose mineralization which they found to be negatively correlated with the concentration of copper and suggested it as a useful indicator of pollution stress. Manley & Davenport (1979) discuss shell closure in marine molluscs in response to increased copper levels. In another study (Harrison & Rice, 1979), the greatest mortality of oysters was found to occur over 96 h at intermediate levels of copper but not at the highest levels, a factor related to the amount of time the shells were open. Stirling (1975) described the toxicity of copper to Tellina tennis in terms of burrowing capacity. Regressions of copper concentration and the burrowing response of Macoma balthica were found to be insignificant (McGreer, 1979a,b). It was found, however, that there was an active avoidance response by M.balthica to sediments containing the highest metal levels. Variability in toxicity levels can be produced as an artifact, by experimental procedure. Physical conditions of toxicity experiments (i.e., salinity, temperature, and oxygen) must be regulated to provide a constant environment. The added stress to the organism, of coping with more than one perturbing condition, influences the effects produced by the test (Erickson, 1972). Holm-Jansen (1948) examined the effect of heavy metals on osmotic regulation in Daphnia magna and noted that, in diluted sea water, copper became more toxic as the salinity decreased. Hunter (1949) showed that copper was non-toxic in sea water in a range of concentrations up to 50 mg·1−1, to the amphipod Marinogammarus marinus. 0.5 mg·1−1 was toxic in distilled water. Schneider (1972) found 10−2 M copper to be acutely toxic to a marine fungus but noted that this level was controlled by salinity. Olson & Harrel (1973) indicated that the toxicity of copper to an estuarine clam increased with a decrease in salinity. Low levels of copper during periods of persistently low salinities and low or high temperatures were shown to produce intolerable stress upon the recruitment of Crassostrea virginica embryos (MacInnes & Calabrese, 1979). Bovee & Sternshein (1978) found that warming of the water above 30° C accentuates the toxicity of heavy metal ions and reduces the effectiveness of “non-toxic counter ions” in alleviating that toxicity. Examinations of temperature, salinity, and copper combinations during simulated uptake of larval, postlarval, and juvenile pinfish, flounder, and mullet in power plant condenser cooling systems revealed a synergistic effect of copper and temperature (Hoss et al., 1974). Eisler (1977a, b) showed seasonal temperature changes to be reflected in LC-0, 50, 100 values for Mya arenaria while Gibson et al. (1975) found little temperature related change in the 96-h LC50 for Pandalus danae. Another important physical variable is time; most tests are conducted under continuous exposure conditions while in the natural environment organisms are often exposed to toxicants in slug-doses due to intermittent discharge or tidal action sweeping an effluent plume back and forth over an area. Davenport (1977) for example,

Oceanography and marine biology

725

showed Mytilus edulis to be unaffected by copper levels during 6-h on: 6-h off exposures. The organism was able to detect the disturbance and close its valves during the exposures thereby avoiding the perturbation. The effect of copper in toxicity tests may be controlled by the rate of uptake rather than the total amount of copper being accumulated (McLusky & Phillips, 1975). Karrick & Gruger (1976) suggest that consideration should be given to the condition, behaviour, and recovery of the test organism, the overall response of the organism prior to death, and the length of time to the first toxic symptoms. An autopsy should be performed, where possible, to determine the cause of death. Cardeilhac, Yosha & Simpson (1979), in exploring mechanisms for acute copper toxicity to marine teleosts, exposed Archosargus probatocephalus to toxic copper concentrations of 8.5 mg Cu2+·1−1 in sea water. The stages of intoxication, based on behaviour, were lethargy, indifference, lack of coordination, inactivity and death. Autopsies showed both congested and smaller kidneys. Microscopic examination of gills showed lamellae to be blunt and thickened, capillaries congested and mucous glands dilated. These, together with other smaller changes indicated intoxication after copper poisoning was caused by potassium toxicity as a result of damage and failure of osmoregulation by gills and kidneys. Eisler & Gardner (1973) found renal and lateral line lesions in all mummichog subjected to copper concentrations of ≥1 mg·1−1 and epithelia lining the oral cavity necrotized by the caustic action of 8 mg Cu2+·1−1. Differences in response also occur between members of the same species of test organism as a result of the technique used as well as the physiological state of the organism. Erickson (1972) noted an inhibition of population growth on Thallasiosira pseudonana under natural conditions, of 0.68–1.14 µg Cu·1−1 while Mandelli (1969) found the same organism to be inhibited at 0.180–0.265 µg Cu·1−1 in synthetic medium, an increase of two orders of magnitude! The difference was presumed to be a decrease in the activity of copper induced by components of the artificial sea water medium used by Mandelli although part may have been due to the nature of the organism (e.g., Murphy & Belastock, 1980). Referring to the components of natural and synthetic test media, Lee (1973) and Gibson et al. (1975) provide lists of chemical factors important to both marine and freshwater toxicity results: oxidation states, solubility, hardness, complexation, ionic strength, type and amount of solids and adsorption to surfaces (colloidal material, detritus, and other particulates), salt ratios, concentration ratios, and organic content. Tests performed by Okazaki (1976) with Crassostrea gigas gave a 96-h TLm of 0.56 mg·1−1 in a flow through system while Fujiya (1960), under static conditions, obtained a 96-h TLm of 1.9 mg·1−1 for the same organism. Variation in response also occurs as a result of physiological conditions (e.g., Murphy, 1978; Murphy & Belastock, 1980). Sea urchin eggs from different females, for example, were found to give differing results suggesting that comparative tests should be carried out using eggs from the same female (Bougis, Corre & Etienne, 1979). Pre-conditioning of test organisms will greatly influence the results. Artemia salina larvae acclimatized for two weeks to 0.1 mg Cu·1−1 showed enhanced tolerance to 1.0 mg Cu·1−1 when compared to unacclimatized larvae. Adults acclimatized for three weeks to 0.1 mg Cu·1−1 showed similar results (Saliba & Ahsanullah, 1973). Bryan (1976) compared the toxicity of copper (as citrate) to tolerant and non-tolerant organisms and showed that the LC50

Interlinking of physical

1200

concentrations in all cases were markedly higher for tolerant organisms. Harrison, Emerson & Rice (1977) found that organisms from open coastal environments were more sensitive than estuarine specimens, and Moraitou-Apostolopoulou (1978) and MoraitouApostolopoulou & Verriopoulous (1979) observed higher tolerance in Acartia clausii populations from polluted coastal areas than those from relatively clean areas. They found the former populations to be more resistant in terms of feeding, fecundity, longevity, and oxygen consumption, showing no effect at levels of copper (0.001 mg·1−1) which caused supression of non-polluted populations. Other biological factors affecting experimental results are the stages of development used, the number of organisms in the test (culture density), and the number of species involved in any one test. There does not, however, appear to be any correlation between the presumed nutritional requirements of a species for copper and its sensitivity to dissolved copper e.g. molluscs and arthropods with haemocyanin (McLeese, 1974). Erickson (1972) and Jensen & Rystad (1976) note the importance of culture density in determining copper toxicity. Steemann Neilsen & Wium-Anderson (1970, 1971) suggest that dense cultures may overcome the inhibiting action of copper, possibly because of the exudates and dead cells produced and the increased surface area which may lead to complexation, adsorption, and precipitation of copper. Dashora & Gupta (1978) found that moderations of the toxic effect of copper sulphate can also be provided by the combined presence of two algal species in a toxicity test (e.g. Scenedesmus obliquues and Selenastrum minutum). Bougis, Corre & Etienne (1979) used sea urchin larvae in tests to assess water quality and found, however, that increases in the number of pluteii per unit volume of sea water produced little change in LD50. The forms and levels of metals are especially important in trying to relate

TABLE XII Levels of copper found to be toxic to selected marine organisms. *from a review paper, not from original

Species

Toxicity value

Comment

Reference

Bacteria, Yeasts, Protozoa Azobacter—bacteria 10−4 M, Inhibited growth CuSO4 5H2O

Den Dorren De Jong, 1971

Marine bacteria

Marine sulphatereducing bacteria

70 µg/ml

Maximum copper concentration at which at least one strain of bacteria could grow

Kurata & Yoshida, 1978

10−5 M and 10−6 M

CuSO4, slightly stimulatory

Jones & Roche, 1966

10−4 M

inhibitory

10−3 g at./l

Inhibits growth and activity

Hata, 1960

Oceanography and marine biology

Marine fungus

0.5×10−4M

Torula utilis—yeast 30 mg/l

727

Zoospore activity suppressed Weak inhibition of growth

Schneider, 1972

medium pH 5 after steriliization

40 mg/1

Growth ceases

30 mg/1

No growth when medium pH 7 after sterilization

Avakyan & Rabotnova, 1966

Phytoplankton and Algae Mixed algal culture 0.25 mg/l

Increased growth

Filip et al., 1979

0.5–0.75 mg/l Decreased growth ≥1.5 µg/l

Lethal

Nannoplankton

>4 µg/l

Photosynthesis “decreased enormously”

Microplankton

>8 µg/l

Photosynthesis “decreased enormously”

Amphidinium carteri

0.075 mg/l

Reduced growth

Braek et al., 1976

Attheye decora

7×10−5 M

I50

Overnell, 1976

Brachiomonas submarina

2–5×10−5

I50

”

Chlorella pyrenoidosa

1µg/l

Considerable influence

Steemann Neilsen et al., 1969

50 µg/l

Inhibits photosynthesis

C.vulgaris

40 µg at./l

Inhibits growth

Den Dorren De Jong, 1965

Coccochloris elebans

30 µg/l

Inhibits growth

Mandelli, 1969

Species

Toxicity value

Comment

Reference

Cricosphaera elongata

10−3 M

Reduced yield

Reish et al., 1978*

Cylindrotheca closterium

1000 µg at./l Reduces photosynthesis 50% 10 µg at./l

Photosynthetic rate decreased by 30%

100 µg-at./l

Occasionally causes clumping of cells and change in shape (fusiform to oval)

Rajendran et al., 1978

Lehman, 1979

Interlinking of physical

1202

Cycotella nana

0.230 µg/ml Inhibits growth

Mandelli, 1969

Dunaliella primolecta

10−3 M

Reduced yield

Reish et al., 1978*

D.salina

>5 mg/1

Lethal

Pace et al., 1977

D.tertiolecta

7×10−5 M

I50

Overnell, 1976

>0.5 µg/ml

Inhibits growth

Mandelli, 1969

50 µg/l

Inhibits growth (natural population at Russell & Morris, Rhosneigr, Anglesey) 1970

500 µg/l

Inhibits growth (ship’s hulls)

Ectocarpus siliculosus

Exuviaella sp.

0.045–0.025 Inhibits growth µg/ml

Mandelli, 1969

Glenodinium foliaceum

0.055–0.03 µg/ml

”

”

Glenodinium sp.

0.055–0.03 µg/ml

”

”

Gonyaulax tamarensis

10−10·4M

50% of cells non-motile; do not divide or grow larger

Anderson & Morel, 1978

10−9.7 M

Calculated cupric ion activity yielding 100%non-motility

0.1 mg/l

In vitro MLD

0.2 mg/l

In vivo MLD

G.splendens

20 µg/1

≥50% decrease in cell numbers by Day 2

Saifullah, 1978

Microcystis aeruginosa

100 µg/1

Inhibits photosynthesis

Myhal’. 1970

Nitzschia closterium

0.190–0.130 Inhibits growth µg/ml

Mandelli. 1969

0.0329 mg/l 96 h EC50

Rosko & Racklin, 1975

N.palea

1.2 µg/l

30% reduction in photosynthesis

Steemann Nielsen & Wium-Anderson, 1971

Phaeodactylum tricornutum

0.4 mg/l

Growth rate reduction

Jensen& Rystad, 1976

0.41 mg/l

Decreases growth rate by 50%

Hannan & Patouillet, 1977

10−3M

Reduced yield

Reish et al., 1978*

Gymnodinium brevis

Starr& Jones, 1957

Oceanography and marine biology

Species

Toxicity value

Polysiphonia sp.

729

Comment

Reference

No. of months exposure before fouling appeared on painted surfaces at Tahiti Beach,

Weiss, 1947

Cu in paint

Florida

90% dry wt

3

67 ”

2

45 ”

2

30 ”

2

23 ”

2

14 ”

2

Procentrum micans

20 µg/l

≥50% decrease in cell numbers by Day 2

Saifullah, 1978

Scenedesmus quadricauda

0.6–32 mg/l

Growth inhibited. 15 days exposure; administered as dichloro dipyridine copper (II)

Phillip, 1973b*

Scrippsiella faeroense

5 µg/l

≥50% decrease in cell numbers by Day 4–5 Saifullah, 1978

10 µg/l

≥50% decrease in cell numbers by Day 3

20 µg/l

≥50% decrease in cell numbers by Day 1

0.05 µg/ml

Growth inhibition occurred

Mandelli, 1969

5×10−5 M

L50

Overnell, 1976

0.01 mg/l

Growth rate reduction

Jensen& Rystad, 1976

Tetraselmis sp.

10−3 M

Reduced yield

Reish et al., 1978*

Thalassiosira fluviatilis

0.180–0.265 Inhibits growth µg/ml

Mandelli, 1969

T.pseudonanu

0.68–6.14 µg/l

”

Erickson, 1972

0.2 mg/l

Lethal effect

Braek et al., 1976

230 µg/l

Inhibits growth

Jensen & Rystad, 1976

3×10−11 M

Growth inhibited

Sunda & Guillard, 1976

Skeletonema costatum

Interlinking of physical

Species

1204

5×10−9 M

Growth ceases

5–30 µg/l

Populations displayed growth inhibition over entire range

Erickson, 1972

Toxicity value

Comment

Reference

0.089 mg/l

24 h LC50

Reeve et al., 1976

0.2 mg/l

28 day LC50, adult

Reish et al., 1976

0.18 mg/l

96 h LC50, trochopore

0.2 mg/1

96 h LC50, adult

Annelida Larvae Polychaeta Capitella capitata

Galeolaria caespitosa 4.5×10−5 M 50% mortaliy in 2 h

Wisely & Blick, 1967

Hydroides parvus

Weiss, 1947

No. of months exposure before fouling appeared on painted surfaces at Tahiti Beach, Cu in paint Florida 90% dry wt None after 4

Neanthes arenanceodentata

N.japonica

67 ”

”4

45 ”

”4

30 ”

4

23 ”

4

14 ”

3

0.044 mg/l

28 days LC50, without sand

0.10 mg/l

28day LC50, with sand

0.3 mg/l

96 h LC50, juvenile

0.14 mg/l

28 day LC50, juvenile

0.03 mg/l

96 h LC50, adult

0.25 mg/l

28 day LC50, adult

0.5 mg/l

Wt gain and food conversion efficiency decreased

1 mg/l

Survival satisfactory

3 mg/l

Zero survival

Pesch& Morgan, 1978

Reish et al., 1976

Inamori & Kurihara, 1979a

Oceanography and marine biology

Nephthys hombergi

731

0.7 mg/l

96 h LC50; tolerant organisms in 100% sea water

0.25 mg/l

96 h LC50; non-tolerant organisms in 100% sea water

Bryan, 1976

Species

Toxicity value

Comment

Nereis diversicolor

1 mg/l

<1 day LT50 in 50% sea water for low-copper tolerance worms; ≥4 days LT50 for higher-copper tolerance worms

Nereis sp.

Reference

2.3 mg/l

96 h LC50; tolerant organisms in 50% sea water

Bryan, 1976

0.54 mg/l

96 h LC50; non-tolerant organisms in 50% sea water

Bryan, 1976

Estuary

Typical sediment Cu mg/l

Worms initial Cu mg/l (dry)

Avon

MLTs as percentage of Avon controls: mg/l 0.5

1.0

2.5 Mean %

18

20

100 100 100 (155h) (59 (27 h) h)

100

Gannel

296

116

128 100 135

121

Tamar

509

397

Hayle Restronguet Creek Ophyrotrocha 0.05 mg/l labronica Spirorbis lamellosa

Bryan & Hummerstone, 1971

97

93 111

100

712

729

490+ 230 260

327

3500

922

440+ 299 259

333

Growth rate significantly decreased

8.0×10−6 M

50% mortality in 2 h

6×10−5 M

50% mortality in 2 h

>0.3 mg/1

Kills larvae; inhibits adult growth

Brown & Ahsanullah, 1971 Wisely & Blick, 1967

Bryozoa Bugula neritina

0.2–0.3 mg/1 <0.2 mg/1

Retards growth and inhibits polyp formation Retards growth and polypide development

Miller, 1946

Interlinking of physical

Watersipora cucullata

9×10−6 M

1206

50% mortality in 2 h

Species

Toxicity value

Comment

Watersipora cucullata

No. of months exposure before fouling appeared on painted surfaces at Miami Cu in paint Beach, Florida

Wisely & Blick, 1967

Reference Weiss, 1947

90% dry wt 9 67% ”

9

45% ”

8

30% ”

2

23% ”

2

14% ”

1

Cnidaria Campanularia flexuosa (hydroid)

0.010– Inhibition of growth rate, 11-day 0.013 mg/l exposure

Stebbing, 1976

1µg/1

Lowest concentration causing lasting inhibition of growth rate

Stebbing & Hiby, 1979

3 mg/l

Tissue disintegration within a few hours

Karbe, 1972

0.06–1 mg/l

Morphological changes and tissue reorganization

0.03–0.06 mg/l

Threshold concentrations for acute effects

2–12 µg/l

Significant inhibition of reproduction rate Stebbing & Pomroy, 1978

0.5–2.5 µg/l

Control mechanism regulates asexual reproduction enabling organism to counteract inhibitory effect of copper

36 µg/l

24 h LC50

Reeve et al., 1976

146 460 µg/l

24 h LC50

Reeve et al., 1976

68

24 h LC50

Eirene viridula (hydroid)

Hydra littoralis (hydroid)

Phialidium sp. (medusae) Chaetognatha Sagitta hispida biomass: (µg carbon)

394 µg/l

Oceanography and marine biology

Species

Toxicity value

Comment

36

315 µg/l

24 h LC50

0.2

43 µg/l

24 h LC50

40 µg/l

Inhibition of the fertilization reaction

733

Reference

Enchinodermata Arbacia sp. Echinometra mathaei

Lillie, 1921 Heslinga, 1976

gametes

0.18 mg/l

50% fertilization success

fertilized eggs

0.42 mg/l

50% cleavage success to 8-cell stage

larvae

0.02 mg/l

Suppression of skeletal development, 90 h

adults

0.30 mg/l

96 h TLm50

Ctenophora Mnemiopsis mccradyi biomass: (µg carbon)

Pleurohrachia pileus

Reeve et al., 1976 2480 29 µg/l

24 h LC50

185

17 µg/l

24 h LC50

33 µg/l

24 h LC50

”

100 µg/l

24 h LC50

”

0.001 mg/l

Ingestion rate decreases (less so for animals from polluted area). Egg production decreases for ‘clean’ animals, increases for ‘polluted’ animals. Oxygen consumption increases for all animals

Moraitou-Apostolopoulou & Verriopoulos, 1979

0.082 mg/l

48 h LC50, pollution

Moraitou-Apostolopoulou,

Rotifera Brachionus plicatilis Crustacea Acartia clausi

Interlinking of physical

1208

adapted population

1978

0.034 mg/l

48 h LC50, population from nonpolluted area

A.simplex

0.20 mg/l

24 h LC50

Arnott & Ahsanullah, 1979

A.tonsa

104–311 µg/l

24 h LC50

Reeve et al., 1976

9.0–78.0 µg/l

72 h LC50

Sosnowski et al., 1979

Species

Toxicity value

Comment

Artemia salina biomass: (µg carbon):

Reference Reeve et al., 1976

0.86 2050 µg/l

24 h LC50

0.76 2554 µg/l

24 h LC50

2 µg/l

Amylase activity disturbed in 24 h; trypsin activity disturbed in 72 h

Alayse-Danet et al., 1979

14 µg/l

48 h HC50 (halving hatching rate)

Jørgensen & Jensen, 1977

0.16 µg/l

Minimum effect level

Balanus amphilrite

No. of months exposure before fouling appeared on painted surfaces at Miami Cu in paint

Beach, Florida

90% dry wt 4 67%

3

45% ”

3

30% ”

2

23% ”

1

14% ”

1 Tahiti Beach

90% dry wt none attached to paint after 4 months of exposure 67% ”

non attached to paint after 4 months of exposure

Weiss, 1947

Oceanography and marine biology

45% ”

4

30% ”

4

23% ”

3

14% ”

1

B.improvisus

735

No of months exposure before fouling appeared on painted surfaces at Cu in paint

Weiss, 1947

Miami Beach, Florida

90% dry wt none after 10

Species

67% ”

none after 10

45% ”

none after 10

Toxicity value

Comment

30% ”

none after 10

23% ”

2

14% ”

1

2778 µg/l

24 h LC50

250 µg/l

LT50 =30.0 days (LT0=20; LT100=68)

500 µg/l

LT50=7.7 days (LT0=3; LT100=4 49)

megalops

500 µg/l

LT50=3.7 days (LT0=1; LT100=3 31)

Cancer irroratus

>5 mg/l

Mortality occurred

Thurberg et al., 1973

adult

109 mg/l

48 h LC50

Connor, 1972

larvae

0.6 mg/l 0.1 meq

Decrease in respiration rate: 34% in heart tissue 28% in gill tissue 6% in hepatopancreas tissue 3% in thoracic ganglia tissue

Kerkut & Munday, 1962

100 meq

all tissues respired at <50% normal rate

90 µg/l

24 h LC50

Calanus plumchrus Callinectes similis 1st crab stage

Reference

Reeve et al., 1976 Neff & Anderson, 1977

Carcinus maenas

Copepod larvae

Reeve et al., 1976

Interlinking of physical

1210

Corophium acherusicum

1.4 mg/l

96 h TLm50

Bellan-Santini & Reish, 1976

C.volutator

50 mg/l

168 h LC50-tolerant organisms

Bryan, 1976

32 mg/l

168 h LC50-non-tolerant organisms

adult

29.5 mg/l

48 h LC50

Connor, 1972

larvae

0.33 mg/l

Daphnia magnum

0.03 mg/l

100% mortality in 24 h in 1:100 ocean water

Holm-Jensen, 1948

Crangon crangon

Species

Toxicity value

Comment

Euphausia pacifica

Reference Reeve et al., 1976

24 h LC50

biomass: (µg carbon) 2957 14 µg/l 30 µg/l

24 h LC50

Gammarids

90 µg/l (1.4 mmol/l)

30 day LC50

Hueck, 1975

Gammarus pseudolimnaeus

1.7 mg/l

96 h LC50

Bellan-Santini & Reish, 1976

Homarus gammarus (larvae)

0.033–0.1 mg/l

48 h LC50

Connor, 1972

Labidocerca scotti

132 µg/l

24 h LC50

Reeve et al., 1976

Marinogammarus marinus

≥50 mg/l

Toxic in sea water

Hunter, 1949

0.5 mg/l

Toxic in distilled water

Metridia pacifica

176 µg/l

24 h LC50

Reeve et al., 1976

Nitocra spinipes

1.8 mg/l

96 h LC50

Bengtsson, 1978

Pandalus dane

Nominal Cu Labile Cu

273

J.S.Young et al., 1979a

5 µg/l

0.66 µg/l

Zoeal development complete, 4th moulting delayed

10 µg/l

0.88 µg/l

Zoeal development complete, 4th moulting delayed

20 µg/l

7.7 µg/l

All zoea died in 1st or 2nd stage

50 µg/l

21.4

All zoea died in 1 st stage

Oceanography and marine biology

737

µg/l 0.04 mg/l

Growth retardation, 1 month exposure

Gibson et al., 1975

10° C

0.037 mg/l

96 h LC50

”

10° C

0.066 mg/1

96 h LC50

”

15° C

0.021 mg/l

96 h LC50

”

15° C

0.049 mg/l

96 h LC50

”

20 C 0.031 mg/l

96 h LC50

”

20° C

0.042 mg/l

96 h LC50

”

Paracalanus parvus

0.19 mg/l

24 h LC50

Arnott & Ahsanullah, 1979

Paracereis sculpta

1.25 mg/l

96 h TLm50

Bellan-Santini & Reish, 1976

Paragrapsus quadridentatus

0.17 mg/l

Larvae, 96 h LC50

Ahsanullah & Arnott, 1978

exposure temp.

Species

Toxicity value

Comment

Reference

Penaeus aztecus and P.douorarum

0.05 mg/l In sea water brine mixtures toxic Mandelli & to nauplius, protozoeal, and mysis McIlhenny, 1971 stages

P.californiensis

250 mg/l 96 h TLm50

Hanks, 1976

Podocerus fulanus

0.32 mg/l 96 h TLm50

Bellan-Santini & Reish, 1976

Scutellidium sp.

0.18 mg/l 24 h LC50

Arnott & Ahsanullah, 1979

Undinula vulgaris

192 µg/l 24 h LC50

Reeve et al., 1976

Pelecypoda Anadara granosa A.senilis Cardium edule

60 µg/l 96 h LC50 0.120±0.123 Interruption of normal behaviour mg/l 95% confidence limits ≈1.0 mg/l 48 h LC50

Kumaraguru & Ramamoorthi, 1978 Manley & Davenport, 1979 Eisler, 1977b*

Interlinking of physical

Crassostrea gigas

1212

650±100 µg/l 48 h LC50

Harrison & Rice, 1979

430±60 µg/l 96 h LC50 230 µg/l Incipient LC50 1.9 mg/l 96 h TLm (static conditions) 0.085±0.025 Interruption of normal behaviour, mg/l 95% confidence limits embryos

20 µg/l 100% mortality in 48 h Cu effect on embryo development

Fujiya, 1960 Manley & Davenport, 1979 Knezovich & Harrison, 1978 Harrison, 1977

Cu2 + % mortality 48 h toxicity test 100 µg/l disintegrated by 24 h 80 µg/l disintegrated by 36 h 40 µg/l 100% 20 µg/l 100% 10 µg/l 45% 5 µg/l 18% control 5% 88 µg/l 96 h LC50

C.madrasensis

Species

Toxicity value

Comment

C.virginica larvae

Kumaraguru & Ramamoorthi, 1978

Reference Mandelli & Mcllhenny, 1971

0.02 mg/l (0.01 ionic Cu)

Desalination brines, toxic effect

0.05 mg/l

LC6 (20 wk)

Eisler, 1977b*

larvae

>0.5 mg/l

Toxic

Prytherch, 1934

embryos

0.103 mg/l

LC50 (48 h)

Calabrese et al., 1977

larvae

0.0328 mg/l

12-day LC50

Mercenaria mercenaria

0.025 mg/l

LC91 (20 wk)

Eisler, 1977b*

larvae

0.0164 mg/l

8–10 day LC50

Calabrese et al., 1977

Meretrix casta

72 µg/l

96 h LC50

Kumaraguru & Ramamoorthi, 1978

Modiolus

≈0.05 mg/l

Interruption of normal behaviour

Manley & Davenport,

Oceanography and marine biology

739

demissus

1979

M.modiolus

0.141±0.108 mg/l Interruption of normal behaviour 95% confidence limits

Mya arenaria

35 µg/l

96 h LC50

Leland et al., 1978*

0.035 mg/l

168 h LC50

Eisler, 1977b

0.039 mg/l

96 h LC50 (LC0=0.025; LC100=0.100)

”

5.000 mg/l

48 h LC50 (LC0=0–150; LC100=15–000)

”

summer (T=22° C)

0.035 mg/l

168 h LC50 (LC0=0–025; LC100=0–050)

”

fall(T=17.5°C)

0.086 mg/l

168 h LC50 (LC0=0.075; LC100=0.100)

”

0.086 mg/l

336 h LC50 (LC0=0.075; LC100=0.100)

”

0.086 mg/l

504 h LC50 (LC0=0.075; LC100=0.100)

”

168 h LC50 (LC0=3.000; LC100=>3.000)

”

>0.02 mg/l

LC100 (several weeks)

”

0.035 mg/l

LC50 (168h) (LC0=0.025; LC100=>3.000)

winter (T=4.0°C) >3.000 mg/l

’

Species

Toxicity value Comment

Reference

Mytilus edulis (Plamulatus larvae)

3.5×10−4M

2 h LD50

Wisely & Blick, 1967

0.012 mg/l

LC0 (30 days)

Eisler, 1977b*

0.025 mg/l

LC50 (19 days)

0.045 mg/l

LC100 (10days)

0.2 mg/l

LC55 (7 days) & toxicity threshold Scott & Major, 1972

3.5×10−4M

50% mortality in 2 h

Wisely & Blick, 1967

0.5 mg/l

MLT=2 days

Davenport, 1977

0.25 mg/l

MLT=4–5 days

500 mg/1 CuNa Inhibits oxygen consumption of citrate whole animal

Brown, 1972

Interlinking of physical

1214

0.3 mg/l

TLm 9–10 days

Delhaye & Cornet, 1975

0.3 mg/l

7 day LC50

Martin et al., 1975

0.28 mg/l

96 h LC50

Abel, 1976

0.26 mg/l

7 day LC50

0.1–0.2 mg/l

7 day TLm

Marks, 1938

3 mg/l

9–10 days TLm

Delhaye & Cornet, 1975

Jan., Feb. (prespawning)

1 mg/l

9–10 days TLm

”

Mar. (beginning spawning)

1 mg/l

6 days TLm

”

2–3 days TLm

”

Apr., May (spawning) 1 mg/l

0.021±0.01 mg/l Interruption of normal behaviour, 95% confidence limits

Manley & Davenport, 1979

Mussels

15 µg/l (0.24 mmol/l)

30 day LC50

Hueck, 1975

Oysters (embryos)

100 µg/l (1.6 µmol/l)

96 h LC50

Hueck, 1975

Japanese oysters

1.9 mg/l

TLm (96h)

Fujiya, 1960

Rangia cuneata

0.78 mg/l

TLm48 (TLm72=0.25 mg/l TLm96=0.21 mg/l) salinity=<1‰

Olsen & Harrel, 1973

Species Rangia cuneata

Tellina tennis Venerupis decussata Villorita cyrinoides var. cochinensis Gastropoda

Toxicity value

Comment

Reference

16.0 TLm48 (TLm72=11.7 mg/l TLm96=8.0 mg/l) salinity=5.5‰

Olson & Harrel, 1973

14.7 TLm48 (TLm72=10.0 mg/l TLm96=7.4 mg/l) salinity=22%0

”

1000 µg/l 96 h LC50 0.1 mg/l LC 100 (50 days) 2 mg/l 240 h LC50

Stirling, 1975 Eisler, 1977b* Lakshmanan & Krishnan Nambisan, 1977

Oceanography and marine biology

741

50 µg/l 96 h LC50

Haliotis cracherodii

H.rufescens

Martin et al., 1977

>32 µg/l Histopathological abnormalities in gill tissue

”

65 µg/l TLm adults 114 µg/l TLm larvae

”

>32 µg/l Histopathological abnormalities in gill tissue Fish Clupea harengus

Blaxter, 1977

eggs

30 µg/l High mortality

larvae

1000 µg/l High mortality

Continuous exposure for 180 h;

Rice & Harrison, 1978b

embryos (12 h after fertilization through hatching)

µg/1±SD Time to median lethal level±95% confidence limits 38.1±9.4 144.9 h±8.3 44.1±11.6 134.4 h±5.2 51.2±9.7 134.8 h±3.0 127.9±34.6 115.4h±2.8 235.5±47.2 98.7h±2.1

Species Continuous exposure to 300 h; larvae (hatching through yolk sac absorption)

Toxicity value µg/l±SD

Comment

Reference

Time to median lethal level±95% confidence limits

1349.0+247.0 41.7h±7.3 1969.0+148.0 23.8h+1.5 2425.4+89.0

20.9 h+2.4

3430+710.0

15.6 h

Pulsed exposures, embryos (62 through 98 h after fertilization)

93.8±6.3 µg/l 38.4 h±2.0 h=LT50

Pulsed exposures, embryos (98 through 136 h after fertilization)

111.9±13.7 µg/l

Engraulis mordax

Rice & Harrison, 1978b

53.9 h±6.1 h=LT50 Rice & Harrison, 1978a

Interlinking of physical

embryos

larvae

Fundulus heleroclitus

1216

200 µg/l

12 h LC50

190 µg/l

Incipient LC50

460 µg/l

12 h LC50

400 µg/l

24 h LC50

370 µg/l

Incipient LC50

≥1 mg/l

Renal and lateral line lesions

8 mg/l

Epithelial lining oral cavity necrotized

8 mg/l

30% dead in 96 h

Oncorhynchus kisutch

Lorz & McPherson, 1976

(smolting juveniles)

74 µg/l

96 h LC50 (Nov.)

60 µg/l

96 h LC50 (May)

O.tshawytscha

Hazel &Meith, 1970

eggs

0.08 mg/l

No effect on hatching success of eyed eggs

fry

0.04 mg/l

Acutely toxic

0.02 mg/l

Inhibited growth, increased mortality

Species

Toxicity value

Comment

Pleuronectes platessa newly hatched larvae Salmo salar

Reference Blaxter, 1977

1000 mg/l

High mortality

300 mg/l

High mortality

48 µg/l

I.L.L. (incipient lethal level) water hardness=20 mg/l CaCO3

Sprague, 1964b

32 µg/1

I.L.L.—water hardness=14 mg/l CaCO3

Sprague & Ramsey, 1965

Lethal, copper naphthenate

Samylin, 1966

>3000 mg/l Young salmon

Eisler & Gardner, 1973

Sprague, 1964b

Oceanography and marine biology

Temp.

743

Metal (µg/l) Copper Zinc LT50 (h) 260 3 12.2

(°C)

pH

15

7.1– 7.5

”

”

140

”

”

78

3 25.2 (20% survival at 139 h)

”

”

55

3 32.4

„

„

49

3 27.4 (20% survival at 160h)

„

„

46

3 0% dead in 190 h

”

”

25

3 20% dead in 164 h (died of fungus on injured eye)

”

”

2 4160 4.6

”

”

2 2260 7.2

”

”

2 1350 11.5

”

”

2 690 25.2 (40% survival at 254 h)

”

”

2 650 24.4 (40% survival at 160h)

”

”

2 510 0% dead in 113h

”

”

2 388 0% dead in 240 h

”

”

2

17

7.9

2 5520 5.0

”

8.2– 8.4

2 4150 11.8

”

8.6– 8.9

Species

2

3 16.2

3 0% dead in 763 h (control experiment)

3360 25% dead in 280 h

Toxicity value

Comment

”

7.0– 7.2

2 1940 10.3

”

8.8– 9.2

2 1920 20% dead in 217 h

”

8.9– 9.2

2 1310 0% dead in 90 h

”

8.9– 9.3

2

271 0% dead in 960 h

Reference

Interlinking of physical

5

7.1– 7.5

2 2890 28.0

”

”

2 886 0% dead in 209 h

15

”

”

”

Tilapia musambica

1218

109 1250 5.3 41 430 11.6 2.5–1.5 mg/l

High mortality rate of eggs in stage of organogenesis

1.0–0.3 mg/l

Retards hatching and greatly affects survival rate of free embryos when they start feeding

0.1– 0.01 mg/l

Increases pulpitation nd causes hypertrophy of blood circulation organs in moving embryo

Voronina & Gorkin, 1978

Ascidacea Ascidia nigra in sea water in desalination plant effluent

Chesher, 1971 150 µg/l

96 h TLm

80 µg/l

96 h TLm

metal activity with biological effect (e.g. Anderson & Morel, 1978). The reader will notice that some authors do not discuss total copper levels in a toxicity test, but only levels of added metal. Metal contamination (e.g. salts in synthetic media) or background levels of metals and other inorganics and organics in natural sea water can alter the toxicity of copper. The nature of the added metal is also important, Floch, Deschiens & LeCorroller (1964) examined the effects on molluscs of a variety of copper compounds. They found that Cu2O was lethal at 2.25 mg·1−1, CuCl2 at 2.5 mg·1−1, and elemental copper at 50 mg·1−1. They also note Cu+ has a greater toxic effect than does Cu2+, which is more toxic than elemental copper. Paulin, Camey & Pereira (1963) found similar results with Cu2O. Saliba & Krzyz (1976) showed the toxicity of copper salts to Mya arenaria to decrease in the order . In examinations of shell closure of three Mytilus sp. Manley & Davenport (1979) were uncertain whether total copper was the major factor stimulating the behaviour or whether it was one or more of the many copper-organic and copper-inorganic complexes. J.S. Young et al. (1979a) and Harrison & Rice (1979) both determined levels of labile as well as total copper and noted distinct differences in biological activity of the forms. Hannan & Patouillet (1979) conducted algal toxicity tests on sediment elutriates and suggested that bulk metal analysis of sediments were not sufficient in assessing the toxicity of dredge spoils. The interaction of toxicants is also of importance (e.g., Karrick & Gruger, 1976; Alabaster, 1976). The combined effect can be beneficial (antagonism) or detrimental

Oceanography and marine biology

745

(synergism) (e.g., Alabaster, 1976; Braek, Jensen & Mohus, 1976). Brown & Dalton (1970) have used mathematical models with good agreement between predicted and actual toxicity values. The inhibitory effect of copper on the colonial hydroid Campanularia flexuosa was modelled by Stebbing & Hiby (1979) in order to assess the counteractive effect of the growth control mechanism. A synergism has been noted between copper and nutrients such as nitrates and phosphates. Meijer (1972) showed a decrease in phosphate concentration to cause an increase in the toxicity of copper to Chlamydomonas sp. Phosphates were also found to prevent copper inhibition of the marine yeast Rhodotorula rubra (Button & Dunker, 1971). The possible production of a complexing agent, rhodotorulic acid, by this organism should also be considered. The addition of chlorine has the potential for modifying the organic compounds that complex copper in natural waters, thus increasing the toxicity of the copper present (Carpenter & Smith, 1978). Eisler & Gardner (1973) noted that the synergistic effects of copper and zinc in estuarine mummichog were related to different individual toxic properties of the metals at different anatomical locations or sites of activity. An inverse correlation (P<0.05) between log LC50 and adult Acartia tonsa density at the time of collection has been described (Sosnowski et al., 1979). CHANGES IN THE EFFECT OF COPPER DURING THE LIFE HISTORY OF THE ORGANISM The toxicity of copper sponsored much of the earlier work with this metal and marine organisms. In addition, during the last part of the nineteenth century and first part of the twentieth, there was a great interest in the embryology and life history of organisms. Studies of the effect of copper on stages in the life history of marine organisms were thus natural. In the developing organism, the response to environmental factors changes. A certain amount of copper and other metals are needed to combine with organics during the formation of essential organo-metallic compounds (e.g., Bowen, 1966). One of the earlier studies dealt with the effect of copper on fertilization and cleavage of echinoderm eggs (Lillie, 1921). Lillie operated from the standpoint that copper salts were injurious to many organisms and that fertilization was the “activation of a substance, contained in the cortex of the egg, by the spermatozoan (Lillie, 1914; 1919)”. He believed that copper might inhibit that activation and affect the process of fertilization. His statement that “if copper chloride inhibits fertilization by combining with the fertilizing (the ‘active substance’ in the egg cortex) of the egg” is one of the first statements of the potentially toxic effects of metals on ontogeny, by modifications of active agents in marine organisms. Lillie’s work (1921) indicated that copper can inhibit fertilization and that certain organics (gum arabic, gelatin) can protect against the effect of copper albeit incompletely. Glaser (1923) found that the eggs of the sea urchin Arbacia will take up copper from sea water, that there is an exchange of copper with the environment, and that certain areas of the egg (chorion primarily) appear to be involved with the exchange of copper. Glaser (1923) also suggested the association of copper with certain enzymes, the apparent need for copper by the organism, the possible interplay between metals both inside and outside

Interlinking of physical

1220

the egg, and demonstrated the effect of copper on cleavage of the egg. Much of Glaser’s work was, however, of a histological nature and required washing in tap water; the effect of metals in the tap water was not considered. Cleavage rate of Arbacia eggs has been found to be increased by small additions of copper chloride 10−7–10−13M), an indication that the metal is needed in small amounts by the developing embryo (Allee et al., 1941; Finkel et al., 1942). Ozoh (1980) incubated zebrafish eggs (Brachydanio rerio) in copper-enriched water with and without the shell membrane, and noted the protection provided by the membrane. He also discussed the embryological abnormalities and deformations caused by the experiment. Weiss (1947, 1948) deals with the abnormal growth and development of fouling organisms with copper and mercury antifouling paints. In life history studies of Crassostrea gigas F.L.Harrison et al. (1977) show increased copper levels to increase mortality, larval abnormalities, oxygen consumption, and heart rate, decrease hatching size and change the behaviour of the developing eggs and larvae. Jørgensen & Jensen (1977) report effects of CuCl2 additions on Artemia salina hatching, at concentrations of the same levels found in natural sea water. In studies of the early ontogenesis of Tilapia Voronina & Gorkin (1978) document the levels of copper required to produce a number of disorders in developing embryos (Table XII). Glaser & Anslow (1949) found, in ascidian tunicates, that copper decreased the length of time spent in the larval form (swimming stage) and caused early metamorphosis into the adult. They also found that copper in water conditioned by the metamorphosing of other ascidians gave the same result. The increase in copper was roughly proportional to the number of metamorphoses that had occurred in the medium. Glaser & Anslow (1949) felt that the shorter larval life was due to the poisoning of larval enzymes. This was primarily because of a statement in work by Grave & Nicoll (1939) (and a similar one in Grave, 1941) that “any material…that poisons [larval] enzymes, either wholly or in part, would decrease the length of larval life and hasten the appearance of metamorphosis”. Glaser & Anslow (1949) also suggested that the copper levels initially “activate” larval enzymes but that, as the larvae begin to metamorphose and copper levels supposedly increase, the effect is to poison the enzymes and cause an increase in the numbers metamorphosing. The effect of dilution in the natural environment (e.g., Riley, 1937; Ireland, 1973) of metabolites (Lucas, 1947, 1949), and of complexing agents (e.g., Provasoli, 1963; Barber & Ryther, 1969; Lewis et al., 1972, 1973) also need to be considered in determining the rôle that copper plays in development. In addition, although the settlement of ascidian larvae from the plankton may be accelerated by the addition of copper, the same amount of copper may be toxic to the settled and metamorphosed animal (e.g., Bougis, 1962a). Different stages in the development of organisms respond differently to similar copper levels. The work of Neff & Anderson (1977) on Callinectis similis demonstrates this variability: Copper (µg·1−1)

Life Stage

LT50 (days)

LT0 (days)

LT100

500

Megalops

3.7

1

31

Oceanography and marine biology

1st crab stage

7.7

747

3

49

Reish et al. (1976) have carried out similar comparisons of two polychaetous annelids: Neanthes arenacoedentata

Capitella capitata

LC50

Adult

Juvenile

Adult

Trochophore

96 h

0.3a

0.3

0.2

0.18

28 day

0.25

0.14

0.2

–

aall

units are mg

Cu·1−1.

Reish et al. (1976) found the adults of these two species were more tolerant than the juveniles. Pyefinch & Mott (1948) showed that the larval stages of the barnacle have different sensitivities to copper, the nauplius stages being much more sensitive than the cyprids. TLm values have also been shown to be relative to sexual maturity in the mollusc Mytilus edulis (Delhaye & Cornet 1975). Rice & Harrison (1978b) note that the embryos of Clupea harengus pallasi show significant mortalities at 35 µgCu·1−1. Ahsanullah & Arnott (1978) found that Paragrapsus quadridentatus larvae had greater sensitivity to copper than adults and calculated “potency ratios” of Cu/Cd (3.1) and Cu/Zn (7.2). Anchovy embryos were shown to be more sensitive than the larvae in studies conducted by Rice & Harrison (1978a). Harrison, Emerson & Rice (1977) and Harrison (1977a) noted that, in general, developing eggs and larvae of Crassostrea gigas are more sensitive than adult organisms. This finding is in agreement with Patin et al. (1978) who noted, in a study of certain species of Caspian Sea and Atlantic crustaceans and fishes, that resistance is dependent upon the stage of development and the duration of the experiments. Work by Bernard & Lane (1963) suggested that the toxicity of the cupric ion to the cyprid or the settling stage of barnacles was due to several interrelated metabolic responses. Prytherch (1931, 1934), in his examination of the factors affecting the settling, metamorphosis, and distribution of Ostrea virginica found that salinity and copper controlled the intensity of setting of the oyster spat. Duration of setting was shortest (optimal) with a salinity of 16–18.6‰ and an environmental copper concentration of 0.05–0.60 mg·1−1. This is a high level; Galtsoff (1932, quoted by McKee & Wolf, 1963) found 0.1–0.5 mg·1−1 to be toxic to some oysters. Prytherch (1934) associated the source of copper with the source of fresh water, the drainage system, and stated that “the oyster larva receives the stimulus for setting through the ingestion of copper in the form of a colloidal precipitate and reacts to its presence after an average latent period of 4 minutes and 20 seconds”. His studies also indicated that copper was toxic to the larva when the concentration was in excess of that found in natural waters. This, together with his belief that copper controls settling and development, led him to state that the “horizontal and vertical distribution of oysters in different coastal regions can be correlated with the copper content, salinity of the water, and the variations in these factors under different hydrographical and tidal conditions”. Korringa (1941), in a parallel study of Ostrea

Interlinking of physical

1222

edulis, took exception to Prytherch’s work on the importance of copper and its association with salinity. Korringa doubted the values of copper, suggesting that they were too high. He also doubted the ingestion of copper oxychloride particles because of his finding that both oysters and mussels rejected these particles as pseudofaeces and then closed their shells and ceased feeding. Korringa also believed that a high concentration of oyster larvae, during a specific portion of the tidal period, was due to the concentrating effect of the currents. He did not deny the necessity of copper in the oyster larva and, citing Voisin (1933) as a source, believed that coastal waters contain this element in sufficient quantities. There are many applications of copper toxicity data, one of which is the development of water quality criteria (Environmental Protection Agency, 1978a,b). Whatever multitude of mathematical adjustments and equations are, however, applied there appears to be a tendency to give little consideration to the changing nature of the organism and to environmental characteristics which may vary widely from one location to the next, drastically affecting the applicability of these standards.

WATER QUALITY CRITERIA—WHAT TO MEASURE? A major portion of the copper in the sea is derived from land sources, therefore, it is the coastal regions which receive much of the metal input. The nature of both natural and anthropogenic sources of copper varies widely (e.g., Wyatt, 1978; Gross et al., 1971) although any accumulation in estuarine regions will be evident in the water, the sediments, or the organisms (e.g., Waldhauer, Matte & Tucker, 1978; Trident Engineering Association, 1977). The toxic effects of copper are, however, controlled by a number of factors (e.g., Templeton, 1978; Hutchinson & Collins, 1978) and the establishment of water quality objectives based on experiments with single metals, without considering effects of other variables has been shown to be inadequate (e.g., Wong, Chau & Luxon, 1978b). Large scale scientific projects such as the Controlled Ecosystem Pollution Experiment (CEPEX) have been designed to examine some of the numerous factors affecting copper toxicity (Menzel & Case, 1977) although even these have been plagued with problems which make resolution of the effect of copper difficult (see Gamble, Davies & Steele, 1977; Koeller & Parsons, 1977). Although specific details exist on the toxicity of copper to humans and their protection (e.g., Sittig, 1976), similarly specific details for water quality criteria do not exist, especially for estuarine and marine waters. The Water Quality Criteria established by the U.S. Environmental Protection Agency (1978a,b) appear as a set of guidelines based on aquatic toxicity data from laboratory tests on individual species, using the geometric mean LC50, calculated from all LC50 data on a pollutant. This value is multiplied by a “sensitivity factor” calculated from the varying sensitivities of different species, to give a final “Acute Value”. A comparable process is used to determine chronic values. In the Environmental Protection Agency publication, “Copper: Ambient water quality criteria” (1978b), the following statement is made: “For copper the criterion to protect saltwater aquatic life as derived using the Guidelines is 0.79 µg·1−1 as a 24-h average and the concentration should not exceed 18 [sic, should be 1.8] µg·1−1 at any time.” As a

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qualifying statement however, “It is recognized that the copper criterion approaches the concentrations of dissolved copper reported for saltwater… There is evidence which indicates that, if copper is complexed by organic compounds in saltwater, the toxicity of the metal can be greatly reduced. As a consequence, it is not necessarily the total concentration of copper in saltwater that determines the toxicity to saltwater organisms but the form of the metal that is the toxic component. It must be emphasized, however, that any addition of copper to saltwater above the criterion could exceed the chelation capacity and render saltwater toxic.” Perhaps the biggest set of factors behind the lack of specific details on water quality criteria in the marine environment is the lack of suitable analytical techniques to rapidly determine in situ biological availability of copper in salt water. Due to the chemical changes that occur shortly after sample collecton and the effect of sample treatment (on metal speciation and biological availability) it is impossible to set realistic levels of biologically available copper. Even with massive amounts of information about a given area (e.g., Science Applications, Inc., 1978a–e) the biological effect can only be inferred. The factors that are responsible for controlling the biological effect of copper include both organism-associated and environmental conditions. The ability of dissolved organic matter to react with copper has been shown (e.g., Lewis & Whitfield, 1974) and regional differences have been noted. Recent work (Kerr & Quinn, 1980) has shown that organic matter from different water masses exhibits different chemical characteristics and some differences in the ability to solubilize copper. This type of information is just beginning to provide enlightenment regarding complexes of naturally occurring organic ligands with trace elements which Jenne & Luoma (1977) believe is “the major limitation in calculating overall trace-element speciation in natural waters”. In addition to the environmental factors that make the use of total dissolved copper levels meaningless for water quality criteria, organism-associated factors produce large variations in response both within and between species of organisms (e.g., Murphy & Belastock, 1980). In addition, responses of organisms to heavy metals are frequently obtained at unrealistically high concentrations as pointed out by Lagerwerff (1975) and in a manner which does not allow consideration of natural environmental conditions. As Klapow & Lewis (1979) point out, “lexicological experiments should be designed that have more direct ecological meaning”. Osteryoung (1972) also points out that developing meaningful water quality criteria needs input from chemists as well as biologists. In the results of a workshop on trace metals and marine production (Feely & Curl, 1979) it was pointed out that there is a need for better analytical procedures as well as improved procedures for determining the effect of specific trace metal species on succession of natural populations of marine phytoplankton. Even if there were a way to determine the “biologically available” copper in marine waters, it would have only limited value as the organism varies in response both from species to species and, within a species, with respect to its development, and its physiology. Sosnowski, Germond & Gentile (1979), for example, point out that the sensitivity of Acartia tonsa to copper varies with population density and food ration; Murphy (1978) found that genetic changes occurred in cultures of two species of diatoms with time. Phillips (1979) found concentrations of trace metals other than copper (zinc, cadmium, lead, iron) in the tissues of Mytilus edulis and Fucus vesiculosus to vary independently from one another within a

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particular regions in the Black Sea and Kattegat areas and points out that “metal profiles from each organism would lead to totally different conclusions concerning the relative pollution of the waters of the Baltic Sea and Kattegat”.

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AUTHOR INDEX References to complete articles are given in heavy type; references to pages given in normal type; references to bibliographical lists given in italics.

Abdullah, M.I., 593, 615, 617; 744 See Ambar, I., 42, 47, 50; 72 See Howe, M.R., 44; 72 Abe, T., 486; 528 Abel, E.F., 431, 432, 435, 439, 449; 463 Abel, K.H. See Gurtisen, J.M., 607; 759 Abel, P.D., 734; 744 Abrahamsen, G., 202; 337 Achuthankutty, C.T. See Nair, S., 106; 123 Ackefors, H. See Dybern, B.I., 201; 340 Ackermann, G.R., 587; 744 Ackley, S.F., 28; 33 Adam, H. See Goldschmid, A., 457; 466 Adams, A. See Greig, R.A., 759 Adams, D.D. See Young, R.J., 613; 789 Adams, S.M. See Wolfe, D.A., 561; 788 Adams, V.D. See Filip, D.S., 580; 756 Adolph, C. See Ramachandran Nair, P.V., 422; 470 Aerts, M. See Boeyé, A., 106; 117 Afran, A., 642; 744 Agadi, V.V., 648, 649, 650, 651, 653, 656; 744 Ahearn, D.G. See Crow, S.A., 107; 119 Ahl, T., 637; 744 Ahlquist, N.C., 134; 190 Ahrens, G., 486; 528 Ahrens, L.H., 607; 744 Ahsanullah, M., 731; 744 See Arnott, G.H., 727, 730, 731; 744 See Brown, B., 564, 725; 749 See Saliba, L.J., 719; 778 Aida, J. See Hori, K., 467 Ainley, D.G. See Follett, W.I., 461; 465 Åkesson, B., 204; 337 Akhmedov, A.M. See Patin, S.A., 775 Alabaster, J.S., 539, 739; 744 Alayse-Danet, A.M., 539, 556, 717, 728; 744 Albert, R., 204; 337

Oceanography and marine biology

797

Alberts, J.J. See Leyden, D.E., 767 Albright, L.J., 91, 109, 558; 116, 744 See Palmer, D.S., 110; 124 Al-Daham, N.K. See Sarker, A.L., 456; 470 Alderdice, D.F., 311; 337 Aleem, M.I.H. See Campbell, N.E.R., 103; 118 Alexander, G.A. See McDermott-Ehrlich, D., 682, 709; 769 Alexander, G.V., 556, 564, 659, 666; 744 See McDermott, D.J., 567; 769 See Young, D.R., 591, 658, 666; 788 Alexander, J.E., 583, 636; 744 See Corcoran, E.F., 582; 751 Alexander, M., 96; 117 Alexander, V. See Horner, R., 94; 121 Ali, S.A., 626; 744 Aliev, A.D. See Kasymov, A.G., 326; 343 Allee, W. See Finkel, A.J., 566; 756 Allee, W.C., 566, 740; 744 Allen, D.A. See Austin, B., 559; 745 Allen, H.E. See Mancy, K.H., 554; 769 See Minear, R.A., 770 Allen, H.L., 92; 117 Allen, L.G., 442; 463 Allen, R.R., 233; 337 Aller, R.C., 283, 322, 177; 337, 190 Almeida, S.S. See Guarino, C.F., 588; 759 Alongi, D.M., 236, 282; 337 Alsterberg, G., 277, 280, 303, 320; 337 Alted, M.D., 422, 423; 463 Amann, H. See Mustaffi, Z., 625; 772 Ambar, I., 42, 43, 47, 50, 53, 58, 59, 60, 61, 64, 68, 69, 70, 72; 72 Amiel, A.J., 587, 642; 744 See Navrot, J., 772 Amin, B.S. See Krishnaswami, S., 193 Amio, M., 375; 415 Ammerman, J.W. See Fuhrman, J.A., 98; 120 Amsden, M.P., 592; 745 Amster, J.L. See Means, J.L., 605; 770 Anadon, E., 509; 528 Anantaraman, S., 409; 415 Anctil, M., 473, 499, 503, 505, 506, 507, 508, 510, 511; 528 See Gariépy, P., 505; 529 Anderlini, V.C., 569, 571, 616, 626, 627, 643, 661, 665, 666, 669, 671, 676, 689, 713, 715; 745 Andersen, R.J., 612; 745 Anderson, D.M., 538, 543, 551, 559, 603, 610, 717, 722; 745 See Morel, N.M.L., 537; 770 Anderson, D.T., 375; 414 Anderson, D.Q. See ZoBell, C.E., 106; 127

Interlinking of physical

1272

Anderson, G.C., 77; 117 Anderson, G.M., 641; 745 Anderson, J.B. See Wheeler, R.B., 588; 787 Anderson, J.G. See Meadows, P.S., 236, 283; 345 Anderson, J.M. See Tsuji, F.I., 534 Anderson, J.W. See Neff, J.M., 564, 729; 772 See Rossi, S.S., 330; 347 Anderson, M.A., 603; 745 Andersen, P.D., 618; 745 Anderson, R.F. See Bacon, M.P., 176, 179, 181, 184, 185, 187; 190 See Deuser, W.G., 154; 191 Anderson, V. See Peyton, T., 585; 775 Anderson, Jr, W.D., 422, 423; 463 Ando, R. See Matsui, Y., 769 See Naruse, Y., 771 Andrássy, I., 202; 337 Andreev, V.V., 571, 713; 745 Andren, A.W. See Elzerman, A.W., 587; 755 Andrews, P., 91, 93; 117 Andriashev, A.P., 1, 5, 28, 31; 33 Anger, K., 321, 323; 337 Angino, E.E. See Long, D.T., 604; 767 Ankar, S., 229, 230, 231, 233, 234, 242, 243, 271; 337 Anonymous, 548, 567, 587, 657; 745 Ansell, A.D., 352, 354, 355, 358, 366, 369, 370, 371, 375, 376, 382, 386, 388, 394, 397 Fig. 2 (Brown) facing p. 354, Fig. 7 (Brown) facing p. 376; 415 See Brown, A.C., 351, 388; 415 See Trevallion, A., 356; 417 Anslow, G.A. See Glaser, O., 538, 566, 740; 744 Aoyama, I., 571; 745 Apley, M.L. See Russell-Hunter, W.D., 381; 418 Appleby, A.G., 281, 283; 337 Apts, C.W. See Gibson, C.I., 564; 758 See Young, J.S., 550; 789 Arai, R., 426; 464 Arcuri, E.J., 551, 625; 745 Arlhac, D.P. See Berland, B.R., 541; 747 Arlt, G., 226, 231, 234, 237, 271, 321, 323; 337 Armstrong, D.E. See Eisenreich, S.J., 615; 754 See Elzerman, A.W., 586, 605; 624 Armstrong, E.F., 621; 745 Armstrong, F.A.J., 138, 581; 190, 745 Armstrong, J.E., 612; 745 Armstrong, J.W., 422, 423; 464 Armstrong, P.B., 587, 641, 642, 643; 745 Arnac, M., 571, 578, 636, 637; 745 Arndt, E.A., 242, 304; 337 Arneson, D.W. See Colin, P.L., 480;

Oceanography and marine biology

799

466 Arnold, D., 540; 745 Arnoldus, M.J.H.L. See Van Der Weijden, C.H., 592; 786 Arnott, G.H., 728, 731; 744 See Ahsanullah, M., 730; 743 Arntz, W.E., 336; 338 Aronson, L.R., 451; 464 Arora, A. See Gupta, A.B., 543; 759 Arrhenius, G., 624; 745 Arruda, L.M., 421, 422, 464 Arthur, D.R. See Birtwell, I.K., 212, 214, 221, 222, 226, 229, 231, 260, 264, 269, 271, 276, 280, 281, 283, 291, 293, 295, 298, 301, 302, 306, 321; 338 See Hunter, J., 208, 212, 214, 215, 221, 223, 228, 230, 231, 276, 291, 296; 342 Artuz, I. See Dickman, M., 103; 119 Arzhanova, N.V., 9; 33 Asakawa, S. See Inoue, A., 763 Asano, H. See Iwai, T., 492; 531 Askew, C. See Williams, P.J.Le B., 91, 93; 127 Asper, V. See Takahashi, K., 171; 195 Aston, R.J., 215, 233, 280, 320, 321, 322, 323; 338 Aston, S.R., 595, 601, 643; 745 Astruc, M. See Mericam, P., 600; 770 Atkins, W.R.G., 635, 637; 745 See Armstrong, F.A.J., 580; 744 Attraw, H.C. See Horvath, G.J., 744; 762 Atwater, T. See Speiss, F.N., 126 Aubert, M., 571, 714; 745 See Gauthier, M.J., 541, 619, 620; 757 See Pesando, D., 558; 775 Auerbach, S., 554; 745 Austin, B., 559; 745 Austin, H.M., 422, 423; 464 Austin, J. See McGrath, M.S., 588, 635; 770 Austin, M.J. See Brinkhurst, R.O., 233; 339 Austin, R.W. See Smith, R.C., 134; 195 See Tyler, J.E., 133; 197 Auty, E.H. See MacKay, N.J., 768 Avakyan, Z.A., 559, 721; 745 Ax, P., 225, 295; 338 See Reise, K., 281; 347 Ax, R. See Ax, P., 225; 338 Axler, R.P. See Kimmel, B.L., 154; 193 Ayling, G.M., 552, 575, 662; 745 See Bloom, H., 632, 635; 747 Azam, F. See Fuhrman, J.A., 98; 120 See Hodson, R E., 88; 120 See Hollibaugh, J.T., 97; 119 See King, K.R., 101; 121

Interlinking of physical

1274

See Thomas, W.H., 783 See Vaccaro, R.F., 558; 785 Babich, H., 614, 617; 746 Babinets, A.E., 615; 746 Babitskaya, V.R., 538; 746 Backlund, H.O., 197, 225, 236, 238, 239, 244, 267, 268, 275, 283, 287, 305, 307, 311, 335; 338 Backus, R.H., 526; 528 Bacon, M.P., 176, 179, 181, 184, 185, 189; 190 See Bishop, J.K.B., 116, 191 See Spencer, D.W., 175, 187; 194 Badcock, J., 519; 528 Bader, H., 141, 147; 190 See Hoigne, J., 538; 762 Bader, R.G., 615, 641; 746 Baffi, F. See Frache, F., 757 Bagge, P., 258, 270, 271, 287, 291, 293, 322, 324; 338 Bagheri, E.A., 220, 221, 223, 226, 228, 232, 233, 234, 264, 326; 338 Baguet, F., 503, 505, 506, 507, 508, 510, 511, 512, 514; 528 See Christophe, B., 505, 510, 511; 529 See Zietz-Nicholas, A.M., 507; 533 Baham, J., 610; 746 Baier, R.E., 108; 117 Bailey, D.F., 399; 415 Bainbridge, A. See Corliss, J.B., 118 Bainbridge, A.L. See Sarmiento, J.L., 195 Baissac, P.de B. See Brown, A.C., 352, 403; 415 Baker, A.C., 1; 33 Baker, A.de C. See Hempel, I., 29; 34 Baker, D.J. See Wearn, R.B., 20, 25; 36 Baker, E.T., 136, 142, 150; 190 See Sternberg, R.W., 194 Baker, E.W. See Palmer, H.D., 618; 774 Baker, H.R. See Brinkhurst, R.O., 208; 339 Baker, J.H. See Kornicker, L.S., 496; 532 Baker, J.T.P., 544, 569; 746 Bal, D.V. See Mutsaddi, K., 438, 442, 457; 469 Balbontin, F. See Marusic, E.T., 468 Baldwin, R.J. See Williams, P.M., 588, 607; 787 Baldwin, T.O. See Morin, J.G., 532 Bale, A.J. See Morris, A.W., 571, 581, 583, 594; 771 Balistrieri, L., 185; 190 Balkas, T.I. See Ramelow, G., 776 Ball, J.W., 604; 746 Ball, N.B. See Baham, J., 610; 746 Ballard, J.A. See Oliff, W.B., 346 Ballard, R. See Speiss, F.N., 126 Ballard, R.D. See Corliss, J.B., 118

Oceanography and marine biology See Francheteau, J., 118, 756 Bally, R., 358, 368; 415 Bandel, K., 375; 415 Bang, S.S., 476; 528 See Baumann, P., 476; 529 Bankston, D.C., 541; 746 Bannister, W.H., 536; 746 Banno, K. See Matsui, Y., 769 Banoub, M.W., 91, 93; 117 Banse, K., 99; 117 Banus, M.D., 580; 746 Baptist, J.P., 714; 746 See Hoss, D.E., 553; 762 Barashkov, G.K., 554; 746 Barbara, A., 636, 638, 682; 746 Barber, A.A. See Johnston, W., 549; 764 Barber, R.T., 560, 578, 601, 604, 605; 746 See Cross, F.A., 575; 751 See Dugdale, R.C., 118 See Packard, T.T., 104; 124 See Smith, W.O., 97; 125 See Sunda, W.G., 603, 620; 783 Barbier, M., 107; 117 Bardet, J., 635; 746 Barelli, M. See Aubert, M., 745 Barham, E.G., 526; 528 Barnard, K.H., 351, 352, 365, 368, 375; 415 Barnes, A.T., 497, 509, 510, 515; 528 See Case, J.F., 523; 529 See Tsuji, F.I., 494; 533 Barnes, H., 565, 635; 746 Barnes, H.L. See Crerar, D.A., 624; 752 Barnes, S. See Pulich, W., 579; 776 Baross, J.A., 109; 117 Barry, M. See Eisler, R., 755 Barry, M.M. See Eisler, R., 755 Barsdate, R.J., 601; 746 Barth, R. See Godoy, O.T., 578; 758 Bartlett, L., 619; 746 Bartmann, W.D. See Blüm, V., 437; 464 Barton, M.G., 427, 430, 451, 457; 464 Bartos, M.J., 587; 746 Bartsch, I., 244, 245; 338 See Zander, C.D., 457; 471 Bartz, R., 134; 190 See Zaneveld, R.J.V., 135; 197 Barvenik, F.W., 91; 117 Bascom, W., 587, 629, 644; 746

801

Interlinking of physical

1276

Bassin, N.J., 133, 134, 142; 190 Bassot, J.-M., 478, 486, 489; 528 See Boisvert, H., 486; 529 Batehelder, J.H. See Sanders, J.G., 542; 778 Bates, J.M. See Foster, R.B., 575, 577; 757 Bath, H., 457; 464 Batley, G.E., 594, 600, 608, 614, 618; 746 See Florence, T.M., 617; 755 See Rendell, P.S., 614; 777 Batoosigh, E.J. See Brinkhurst, R.O., 201; 339 Battaglia, B., 312; 338 Batzli, G.O., 422; 464 Baudo, R., 592; 746 Baumann, L. See Baumann, P., 476; 528 Baumann, P., 475, 476; 528 See Bang, S.S., 476; 529 See Jensen, M.J., 531 See Reichelt, J.L., 475, 476; 532 Bavor, H.J. See Perry, G.J., 89; 124 See Volkman, J.K., 126 Bawden, C.A., 592; 746 See Brown, D.A., 749 See Parsons, T.R., 612; 774 Beardmore, J.A. See Battaglia, B., 312; 338 Beardsley, G.R. See Carder, K.L., 150; 191 Beasley, J. See Peden, J.D., 578; 775 Beasley, T.M., 185; 191 See Cherry, R.W., 185; 191 Beattie, J.M. See Lamb, C.A., 536; 767 Bebbington, G.N., 693, 694, 702, 705, 708, 711; 746 Beckman, L., 586; 746 Beebe, W.M., 526; 528 Beer, R.M., 160; 191 Beers, J.R., 558, 713; 746 Beeton, A.M. See Fisher, J.A., 280; 341 Beklemishev, K.V., 23, 26, 29; 33 Belamy, D.J. See Sheppard, C.R.C., 687; 780 Belastock, R.A. See Murphy, L.S., 548, 560, 565, 600, 719, 743; 772 Bel’cheva, T.A. See Saenko, G.N., 581; 778 Belderson, R.H. See Kenyon, N.H., 63; 73 Bell, F.W. See Dow, R.L., 544; 753 Bell, J. See Hanson, C.H., 623; 760 Bell, S.S., 243, 282, 336; 338 See Coull, B.C., 336; 339 Bell, W.H., 101; 117 Bella, D.A. See Ramm, A.E., 105; 125 Bellan, G., 328; 338 Bellan-Santini, D., 730; 746

Oceanography and marine biology Bellinger, E.G., 590; 747 Benayoun, G. See Polikarpov, G.G., 565; 776 Bender, J.A., 624, 713; 747 Bender, M.E., 570, 575, 578, 601, 659, 715; 747 See Haven, D.S., 342 See Huggett, R.J., 593; 762 Bender, M.L., 105, 538, 614, 636; 117, 747 Benedict, A.H., 592; 747 Bengtsson, B.E., 730; 747 Benham, B.R. See Bellinger, E.G., 590; 747 Bennett, E.B. See Wyrtki, K., 9, 14; 36 Bennett, E.O. See Tornabene, T.G., 90; 126 Benon, P., 636; 747 Bentley, O.G. See Lamb, C.A., 536; 767 Bentley-Mowat, J.A., 541, 560; 747 Ben-Yaakov, S., 274; 338 Berdar, A. See Calapaj, R., 750 Bere, A., 611; 747 Berg, K., 233; 338 Berg, U. See Christensen, B., 217, 272, 312; 339 Berge, J.A., 243, 461; 338, 464 Berger, L.R. See Mohankumar, K.C., 109; 123 See Pope, D.H., 110; 124 See Schen, J.C., 110; 124 Berger, P.S., 612; 747 Berger, W. See Dymond, J., 191 Berger, W.H., 165; 191 Berghuis, E.M. See De Wilde, P.A.W.J., 427; 465 Bergman, S.C., 593, 611; 747 Bergmann, H., 593; 747 Berland, B.R., 554; 747 Berman, T., 77, 97; 117 See Pollingher, U., 92; 124 See Williams, P.J. Le B., 90, 94; 127 Bern, H.A. See Fryer, J.N., 429; 465 See Marshall, W.S., 429; 468 See Owens, A., 429; 469 Bernal, P. See Ebeling, A.W., 433; 465 Bernard, F.J., 538, 545, 565; 747 Bernard, F.R., 381; 415 Bernard, P. See Gauthier, M.J., 542, 620; 758 Berner, R.A., 105; 117 See Martens, C.S., 106; 122 Bernhard, M., 629, 640, 644; 747 Berrisford, C.D. See Oliff, W.B., 346 Berry, A.J., 422, 435, 457; 464 Bertelsen, E., 488, 489, 499, 511; 528 See Munk, O., 489; 532

803

Interlinking of physical

1278

Bertholf, L.M., 538; 747 Bertine, K.K., 587; 747 See Kharkar, D.P., 617; 764 Beryukova, T.A. See Saenko, G.N., 580; 778 Bethoux, J.P., 61; 72 Betz, K.-H. See Tiemann, H., 202; 349 Betzer, P.R., 133, 141; 191 See Lerman, A., 142, 151, 168; 193 Betzer, S.B., 675, 689; 747 Bewers, J.M., 595, 617; 747 See Yeats, P.A., 747; 788 Bezborodov, A.A., 601; 747 Bezdek, H.F., 108; 117 Bhalla, P.R. See Pieterse, A., 560; 775 Bhan, S., 428; 464 Bharathi, P.A.L. See Nair, S., 106; 123 Bhat, S.G., 177; 190 Bhatti, M.N. See Sarker, A.L., 457; 470 Bhavnagary, H.M. See Kappana, A.N., 764 Bhosle, N.B. See Agadi, V.V., 648; 744 Bianchi, A.J.M., 106; 117 Biddappa, C.C. See Upadya, G.S., 618; 785 Bienati, N.L., 26; 33 Bienfang, P.K., 81; 117 Biesinger, K.E., 564; 747 Biggley, W.H. See Swift, E., 522; 533 Biggs, R.B., 633; 747 Biggs, R.B. See Bopp, III, F., 643; 750 See Strom, R.N., 750; 782 Bilio, M., 245; 338 See Barbare, A., 747 Billen, G., 91, 93, 94, 96, 103, 105; 117 See Vanderborght, J.-P., 103, 104; 126 Birtwell, I.K., 212, 214, 221, 222, 226, 229, 231, 260, 264, 269, 271, 276, 280, 281, 283, 291, 293, 295, 298, 301, 302, 306, 321; 338 Biscaye, P.E., 134, 136; 191 See Brewer, P.G., 191 Bischoff, J.L., 625; 747 Bishop, J.K.B., 82, 83, 84, 130, 142, 148, 150, 160, 162, 166, 185; 117, 191 Bisson, M. See Tessier, A., 600; 784 Bittel, R. See Aubert, M., 745 Black, G.A.P., 558, 716; 747 Black, J.A. See Gross, M.G., 759 Black, W.A.P., 576, 635; 747 Blackburn, T.H. See Oren, A., 105; 124 Blake, N.J. See Doyle, L.J., 584; 754 Blanc, F. See Benon, P., 747 Blanchard, A.S. See Goulder, R., 554, 559; 759

Oceanography and marine biology Blanchard, D.C., 108, 586; 117, 747 Blaxter, J.H.S., 547, 569, 735, 737; 747 Blick, R.A.P. See Wisely, B., 552, 562, 724, 725, 734; 788 Bliss, C.I., 619; 747 Bloom, H., 633, 636; 747 Blount, R.F., 203; 338 Blüm, V., 428, 436, 439; 463 See Laumen, J., 439; 466 Blumer, L.S., 441; 464 Blumer, M. See Cooper, W.J., 88; 118 Blutstein, H., 605, 636, 637; 748 Boaden, P.J.S., 236, 257, 259; 338 Bobbie, R.J., 90; 117 See White, D.C., 126 Bodansky, M. See Rose, W.C., 536, 647; 778 Boden, B.P., 473, 518, 523; 529 Bodoy, A., 309; 338 Boesch, D.F., 272, 274, 323; 338 See Haven, D.S., 342 Boetus, I. See Gordon, M.S., 465 Boeyé, A., 106; 117 Bogdanova, N.N. See Khristoforoya, N. K., 659; 765 Bohlen, W.F. See Dehlinger, P., 753 Bohn, A., 648, 653, 688, 702, 712, 715; 748 See Elderfield, H., 754 Boisseau, D. See Goering, J.J., 541; 758 Boisvert, H., 486; 529 Bojanowski, R., 605, 609; 748 Bok, C.S., 650, 656, 675, 678, 684, 693; 748 Boke, K. See Muller, R. 429; 469 Boldt, W., 296; 338 Bölter, M. See Liebezeit, G., 76; 122 Bonatti, E., 614, 642; 748 Bond, G.C. See Manheim, F.T., 134, 138; 194 Bone, Q. See Mackie, G.O., 490; 532 Benin, D.J. See Berland, B.R., 540, 554; 747 Boon, J.J., 89; 117 Boone, W.R., 537; 748 Booth, C.R. See Holm-Hansen, O., 88; 120 See Williams, P.M.,107; 127 Boothe, P.N., 716; 748 Bopp, III, F., 595, 643; 748 Bopp, R.F. See Olsen, C.R., 776 Borgmann, U. See Hodson, P.V., 539; 761 Bormann, F.H. See Eaton, J.S., 162; 192 Bostrom, K. See Moore, C., 644; 771 Boto, K.G. See Olafson, R.W., 574; 776 Bougault, H. See Francheteau, J., 119, 757

805

Interlinking of physical

1280

Bougis, P., 557, 566, 719; 748 Boulton, A., 566; 748 Bouma, A.H., 627; 748 See Bassin, N.J., 133; 190 Bouquegneau, J.M. See Noel-Lambot, F., 574; 773 Bouquiaux, J., 629, 642; 748 Bourgade, B. See Benon, P., 747 Bourget, E. See Cossa, D., 67 Bouwman, L.A. See De Jong, V.N., 201; 340 Bovee, E.C., 718; 748 Bowen, H.J.M., 536, 581, 603, 740; 748 Bowen, V.T., 176, 581; 191, 748 See Bankston, D.C., 747 See Goldberg, E.D., 758 See Labeyrie, L.D., 179; 193 See Nicholls, G.D., 574; 772 See Noskin, V.E., 176, 191; 194 Bowes, P.M. See Carlucci, A.F., 103; 118 Bowser, C.J., 624; 748 Boyce, F.M., 57; 72 Boyd, C.M., 100, 141; 117, 191 Boyd, S.H. See Wiebe, P.H., 81, 82, 166; 127, 195 Boyden, C.R., 552, 634, 638, 643, 654, 662, 666, 669, 678, 683, 686, 689; 748 See Daniel, M.J., 426; 464 Boyle, E., 538, 628, 636; 748 Boyle, E.A., 585, 604; 748 See Sholkovitz, E.R., 593; 780 Boylen, C.W., 102; 117 Boyne, J.V. See Mathews, T.D., 659; 769 Bøyum, G., 63; 72 Bradford, W.L., 596, 614, 644; 748 See Carpenter, J.H., 585; 750 Braek, G.S., 551, 619, 721, 723, 739; 748 Branch, G.M., 389; 415 Brand, L.E. See Murphy, L.S., 548; 772 Brandhorst, W., 103; 117 Brandt, E., 316; 338 Brass, G.W., 585; 748 Brauer, A., 516; 529 Brauner, P.A. See Stumm, W., 600; 783 Braunstein, H.M., 539; 748 Brazda, F.G., 318; 338 Breder, Jr, C.M., 421; 463 Breed, R.S., 475; 529 Breger, I.A., 607; 748 Bregnballe, F., 229, 242; 338 Brenner, S. See Simoneit, B.R., 780 Bresk, B. See Eick, K., 204; 341

Oceanography and marine biology

807

Brewer, P.G., 137, 156; 191 See Bacon, M.P., 185; 190 See Balistreri, L., 185; 190 See Gardner, W.D., 192 See Spencer, D.W., 84, 175, 187, 561, 637; 126, 194, 781 Brewster, F.E. See Davis, R.B., 281; 340 Briand, F., 611; 749 Bricker, O.P., 614; 749 Briese, L.A. See Giesy, Jr, J.P., 749; 758 Briggs, K.B., 238; 338 Briggs, S.R. See Summerhayes, C.P., 783 Bright, T.J. See Rezak, R., 777 See Taylor, D.D., 697, 701; 783 Brillet, C., 439, 441, 448, 452, 453, 459; 464 Brinkhurst, R.O., 197, 198, 200, 201, 208, 212, 214, 216, 222, 225, 227, 229, 230, 233, 237, 239, 240, 244, 257, 258, 269, 280, 283, 290, 291, 293, 320, 321, 322, 336, 338; 339 See Appleby, A.G., 280, 282; 337 See Chapman, P.M., 208, 291, 298, 308; 339 See Cook, D.G., 208; 339 See Chua, K.E., 231, 238, 240, 282; 339 See Johnson, M.G., 204, 230, 231, 233; 343 See Wavre, M., 236, 238, 240, 282; 349 Brinkman, F.G. See Spronk, N., 563; 782 Brizarry, N. See Phelps, D.K., 577; 775 Broada, E. See Schuster, I., 583; 779 Brock, M.L., 86; 117 See Brock, T.D., 86; 117 Brock, T.D., 86; 117 See Brock, M.L., 86; 117 See Munro, A.L.S., 92; 123 Brockamp, O., 624; 749 Brockmann, U. See Ittekkot, V., 76; 121 Brockmann, U.H. See Kattner, G.G., 107; 121 Brodie, J.W., 1; 33 Broecker, W.S., 176, 178; 191 See Peng, T.H., 84; 124 See Sarmiento, J.L., 194 Brooks, P.W., 89; 117 Brooks, R.R., 579, 580, 636, 666, 670; 749 Brothers, E.B., 421, 425, 440, 443, 445, 452; 464 Brown, A.C., 351–418; 351, 352, 355, 357, 358, 362, 365, 366, 367, 368, 369, 370, 371, 372, 374, 375, 376, 379, 380, 381, 382, 385, 386, 388, 389, 390, 391, 392, 393, 394, 396, 397, 399, 400, 403, 406, 407, 408, 409, 410, 717, Fig. 4 (Brown) facing p. 358, Fig. vii (Brown) facing p. 382; 415, 749 See Branch, G.M., 388; 415 See Cuthbert, K.C., 354, 403; 416 See Golombick, T., 354, 403, 406; 416 See Krijgsman, B.J., 371; 416

Interlinking of physical

1282

See Meredith, F.L., 354, 365; 417 See Newell, P.F., 354, 399, 400, 402; 417 See Trueman, E.R., 354, 379, 380, 382, 385, 386, 387, 388, 390, 392, 396; 417 Brown, B., 564, 725; 749 Brown, B.E., 563, 734; 749 See Stevens, J.D., 692, 694, 699, 704, 706, 708, 711; 782 Brown, C.M. See Miller, D., 76; 123 Brown, D. See Parsons, T.R., 550, 613; 775 Brown, D.A., 749 Brown, D.H., 548; 749 Brown, F.A., 487; 529 Brown, G.G. See Clapper, D.L., 538; 751 Brown, I.F. See Liebezeit, G., 76; 122 Brown, J.C., 141; 191 Brown, R.J. See Brown, A.C., 352, 408, 409; 415 Brown, S.C., 365; 415 Brown, T.J., 100; 118 See Sibert, J., 90, 92; 125 Brown, V.M., 717, 739; 749 Brown, W.P. See Marquis, R.E., 109; 122 Browne, R.A., 376; 416 Bruchhausen, P. See Eittreim, S., 192 Bruemmer, G. See Lichtfuss, R., 587, 597; 767 Brugmann, L., 645, 647, 681; 749 Bruland, K.W., 159, 178, 636; 191, 749 See Franks, R., 635, 638; 756 See Knauer, G.A., 81, 158; 122, 193 See Moore, W.S., 185; 193 See Soutar, A., 194 Brun-Cottan, J.C., 141, 150, 166; 191 See Lambert, C.E., 193 Brunel, S. See Anctil, M., 506; 528 Brunskill, G.J. See Wagemann, R., 592 786 Bryan, G.W., 539, 548, 562, 563, 570575, 578, 582, 588, 643, 648 655, 658, 659, 661, 665, 670, 671, 682 683, 684, 689, 720, 725, 730; 749 See Gibbs, P.E., 536; 757 See Luoma, S.N., 582, 620, 658; 767 Bryant, M.P., 105; 118 See McInerney, M.J., 106; 123 Bryden, M.M. See Griffiths, D., 31; 34 Brzezinska, A., 593; 749 Buat-Menard, P., 586; 749 Buat-Ménard, P. See Marty, J.C., 123 Bubela, B. See Mooney, J.R., 618; 771 Buch, J.B., 411; 416 Buch, K., 635; 749 Buchanan, A.S. See Canterford, G.S., 582 750 Buchanan, J.B., 352; 416

Oceanography and marine biology

809

See Gauld, D.T., 351; 416 Buchanan, M.J., 586; 749 Buck, J. See Case, J.F., 529 Buck, J.B., 492, 498, 515, 525, 527; 529 Buck, K.R. See Ackley, S.F., 28; 33 Buckley, D.E. See Cranston, R.E., 162; 192 Buckley, J.L., 436; 464 Budd, W.F., 2; 33 Buffle, J., 609; 749 Buikema, Jr, A.L. See Cairns, J., 551; 750 Buinitsky, V.K.L., 1, 28; 33 Bulich, A.A., 554; 749 Bulleid, N.C. See Smith, D.F., 126 Bullock, T.H., 402, 403; 416 Bulnheim, H.P. See Siebers, D., 240; 348 Bülow, T. von, 197, 235, 236, 239, 257, 258, 263, 264, 269, 271, 280, 283, 287, 291, 323, 326; 339 Bultman, J.D. See Southwell, C.R., 622; 781 Bunke, D., 209; 339 See Westheide, W., 204, 208; 349 Bunt, J.S., 1, 27, 30; 33 Burgess, T.J., 424, 443, 445, 447, 449; 464 Burke, S. See Seitzinger, S., 105; 125 Burkholder, P.R., 1, 27, 28; 33 Burlingame, A.L. See Boon, J.J., 89; 117 See Eberhard, A., 529 Burnett, W.C., 613, 643; 749 Burney, C.M., 76; 118 See Sieburth, J.McN., 97; 125 Burnison, B.K., 612; 749 Burns, J. See Case, J.F., 529 Burns, K.A., 332; 339 Burns, N.M. See Hargrave, B.T., 154, 158, 160; 192 Burrell, D.C., 570, 579, 595, 614, 636, 637, 658; 749 See Heggie, D.T., 594, 635, 640, 641; 762 Burton, D.T., 569; 750 Burton, J.D., 593, 594, 600, 617, 629, 640; 750 See Moore, R.M., 750, 634; 770 Burzell, L.A. See Harvey, G.W., 107; 120 Buschbom, R.L. See Young, J.S., 789 Bush, L.F., 271; 339 Busschots, P.M.C.F. See Davenport, J., 427; 465 Bussing, W.A. ,393; 464 Butot, L.J. M., 411; 620 Button, D.K., 550, 551, 559, 739; 750 Cain, R.L., 422, 423, 449; 464 Caine.E.A., 367; 416 Cairns, J., 551, 561; 750

Interlinking of physical

1284

See Burton, D.T., 568; 749 See Ruthven, J.A., 561, 714; 778 Calabrese, A., 733; 750 See MacInnes, J.R., 544, 546, 551, 562, 718; 767 Calapaj, R., 684, 694, 699, 701, 712; 750 Calcott, P.H., 101; 118 Calder, J.A. See Johnson, R.W., 88; 121 Calderwood, M.W. See Cardeilhac, P.T., 750 Caldwell, R.S., 280; 339 Callahan, R.A., 653, 655, 660, 664, 668, 673, 675, 677, 680, 684, 686, 688; 750 Calvert, S.E., 131; 191 See Duchart, P., 191; 753 See Gaskell, S.J., 88; 119 See Morris, R.J., 78, 88, 608; 123, 770 See Skei, J.M., 587; 780 Calvert, T.J. See Calcott, P.H., 101; 118 Calvin, J. See Ricketts, E.F., 420; 470 Cambon, P. See Francheteau, J., 119, 757 Cambray, R.S., 585; 750 See Peirson, D.H., 585; 775 Cameron, A.J. See Rendell, P.S., 615; 777 Camey, T. See Paulin, E., 563, 739; 775 Campbell, C.M., 431; 464 Campbell, L.L. See Bryant, M.P., 105; 118 Campbell, N.E.R., 103; 118 Campbell, P.G.C. See Tessier, A., 600; 784 Campbell, R. See Smith, D.F., 126 Canterford, G.S., 582; 750 Capelli, R., 668; 750 See Viarengo, A., 785 Capobianco, J. See Mudroch, A., 576, 579; 771 Cardeilhac, P.T., 545, 719; 750 Carder, K.L., 150; 191 See Betzer, P.R., 133; 190 See Lerman, A., 142, 151, 168; 193 Cardoso, J., 89; 118 Carefoot, T., 420; 464 Carey, C.L. See Waksman, S.A., 559; 786 Caris, N. See Winner, R.W., 557; 788 Carlberg, J.M. See Ford, R.F., 756 Carlucci, A.F., 87, 103, 104; 117 See Bezdek, H.F., 108; 116 See Williams, P.M., 87, 106, 110, 111; 127 Carmack, E.C., 29; 33 See Foster, T.D., 26, 29; 33 Carney, J.F., 91, 95; 118 Caron, D.A. See Sieburth, J. McN., 126 Carpenter, J.H., 554, 585, 739; 750

Oceanography and marine biology Carr, R.S. See Reish, D.J., 562; 777 Carranza, A. See Francheteau, J., 119, 757 See Speiss, F.N., 125 Carriker, M.R., 403; 416 Carritt, D.E., 614, 617; 750 Carruthers, A.B. See Hollibaugh, J.T., 96; 120 Carson, W.G. See Zitko, P., 606; 789 Carson, W.V. See Zitko, P., 606; 789 Carter, C.C. See Hestand, R.S., 541; 761 Carty, C., 107; 118 Casabianca, M.L. de, 444, 456; 464 Case, J. See Menzel, D.W., 592, 743; 770 Case, J.F., 498, 503, 507, 510, 514, 523; 528, 529 See Anctil, M., 507, 510; 529 See Baguet, F., 503, 505, 507; 529 See Barnes, A.T., 495, 508, 509, 510, 515; 529 See Tsuji, F.I., 494; 533 See Warner, J.A., 495, 507, 523; 533 Caspers, H., 294; 336 Castagna, A., 668; 750 Castelli, V.J., 622; 750 Castilla, J.C., 539, 626; 750 Catalano, E. See De Leo, G., 440; 465 Cataudella, S., 426; 464 Cattell, F.C.R., 585, 634; 750 Caullery, M., 254; 339 Cave, W.R., 550, 636, 637; 750 See Lewis, A.G., 535–789 Cawse, P.A. See Peirson, D.H., 585; 775 Cawthorne, D.F. See Davenport, J., 427; 465 Cernohorsky, W.O., 351; 416 Chabert, D. See Vicente, N., 570; 786 Chacos, N. See Werringloer, J., 539; 787 Chadwick, E.M.P., 443; 464 Chaisemartin, C., 556, 565, 571; 750 Chamberlain, N.A. See Andersen, Jr, W. D., 463 Chamberlin, J.L. See Radwin, G.E., 376; 417 Chan, K.C. See Wong, M.H., 552; 788 Chan, W.-Y. See Swader, J.A., 551, 603; 783 Chandler, G.T. See Lindquist, D.G., 453, 457; 468 Chandra Sekhar Reddy, A. See Subba Rao, B.V.S.S.R., 348 Chaney, R.L. See Owens, L.D., 348; 774 Chang, K.-H., 422, 423, 449, 456; 464 See Lee, S.-C, 441, 456; 468 Chang, P.S., 658, 670; 750 Chanut, J.P. See Cossa, D., 752 Chapman, G.A. See McCrady, J.K., 608; 769 Chapman, J.W. See Anderlini, V.C., 745

811

Interlinking of physical

1286

Chapman, P.M., 208, 291, 298, 309, 322; 339 Charlou, J.L. See Alayse-Danet, A.M., 539; 744 Charlson, R.J. See Ahlquist, N.C., 134; 190 Chase, R.R.P., 166, 169; 191 Chassé, C., 332; 339 Chatel, K.W. See Brown, D.A., 749 Chatelain, R. See Boisvert, H., 486; 529 Chau, Y.K., 602; 750 See Van den Berg, C.M.G., 550; 785 See Wong, P.T.S., 743; 788 Chen, C.P. See Chang, K.-H., 422, 457; 464 Chen, H. See Gardner, L.R., 588; 757 Chen, J.C., 641; 750 Chen, K.Y., 627; 750 See Gupta, S.K., 614, 641, 643; 758 See Lu, J.C.S., 613, 643; 767 Cheng, T.C., 622; 751 Cherian, M.G., 550; 751 Cherni, N.E. See Kriss, A.E., 109; 122 Chernyshev, N.S, See Patin, S.A., 775 Cherry, D.S., 577; 751 See Guthrie, R.K., 555; 758 See Rodgers, J.H., 561; 777 Cherry, R.D. See Beasley, T.M., 191 See Heyraud, M., 185; 192 See Shannon, L.V., 194 Cherry, R.W., 176, 185; 191 Chesapeake Research Consortium, Inc. Baltimore, Md, 576, 580, 597; 750 Chesher, R.H., 552, 589, 631, 738; 751 Chesselet, R. See Buat-Menard, P., 586 749 See Dehairs, F., 142; 191 See Lambert, C.E., 193 See Marty, J.C., 122 Chester, A.J. See Larrance, J.D., 159; 194 Chester, R., 635, 638, 643, 644; 751 See Elderfield, H., 754 Chet, I., 562; 751 Chew, K.K. See Armstrong, J.W., 422; 464 Cheyne, A.R. See George, S.G., 758 Chiba, A. See Honma, Y., 438; 467 Chinnayya, B., 545; 751 Cho, K.Y. See Graham, P.H., 484; 530 Choi, C.I., 77; 118 Choukroune, P. See Francheteau, J., 119, 757 Chow, T.J., 636, 637; 751 Choy, C.K. See Wong, M.H., 552; 788 Christeller, J.T., 84; 118 Christensen, B., 204, 212, 217, 273, 312; 339

Oceanography and marine biology Christensen, C.L. See Schmidt, D.J., 554; 779 Christensen, E.R., 542, 588, 631; 751 Christensen, G.M. See Biesinger, K.E., 564; 747 See McKim, J.M., 568; 769 Christensen, J.P., 111, 114; 118 Christensen, M.S., 457, 459; 464 Christian, R.R., 109; 118 Christie, N.D., 358; 416 Christophe, B., 505, 511; 529 See Baguet, F., 510; 529 Chu, K.Y., 550; 751 Chua, K.E., 232, 237, 240, 283; 339 See Brinkhurst, R.O., 201, 233, 236, 238, 239, 279; 338 Chuiko, V.T. See Varenko, N.I., 570, 714; 786 Chumakova, R.I. See Filimonov, V.S., 511; 530 Chung, Y. See Krishnaswami, S., 193 Church, T.M. See Pellenbarg, R.E., 586, 605; 775 Churchland, L.M. See Chapman, P.M., 322; 339 Chvojka, R. See Bebbington, G.N., 746 Cintron, N.A. See Wood, E.D., 643; 788 Ciraolo, L. See Calapaj, R., 750 Civitella, M.V. See Cataudella, S., 426; 464 Claparède, E., 245, 247, 248, 252; 339 Clapp Laboratories, 622; 751 Clapper, D.L., 538; 751 Clare, J. See Jones, D., 422, 431; 467 Clark, E., 480; 529 Clarke, G.L., 509, 522; 529 Clarke, W.D., 515, 523; 529 Clarke, W.D. See Backus, R.H., 528 Clauson, M. See Dymond, J., 192 Clavell, Jr, C. See Zirino, A., 637; 789 Clavell, C. See Zirino, A.H., 597, 613; 789 Claypool, G., 106; 118 Cleave, C.D.van, 209; 339 Clendenning, K.A., 540, 561; 751 Clifford, C.H. See Smith, K.L., 126 Cline, J.T., 602, 614; 751 Coates, K., 324; 339 Cobet, A.B. See Jones, G.E., 558; 764 Cobler, R., 83, 156; 118, 191 See Dymond, J., 191 Cochran, J.K., 179; 191 See Aller, R.C., 176; 190 Codispoti, L.A., 104, 105; 118 See Pak, H., 104; 124 Coffin, R. See Grossman, G.D., 457, 458; 467 Cohen, D.M., 487; 529

813

Interlinking of physical

1288

See Marshall, N.B., 487; 531 Cohn, D.H. See Leisman, G., 483; 532 See Nealson, K.H., 532 Cole, F.A. See Swartz, R.C., 628; 783 Colebrook, J.M., 78; 118 Coleman, A.W., 86; 118 Coleman, R.D., 570; 751 Coleman, R.L. See Coleman, R.D., 570; 751 Coler, R.A., 237; 340 Coles, G.C., 320; 340 Colin, P.L., 480; 529 Collier, R.S. See Thurberg, F.P., 545; 785 Collier, R.W. See Bishop, J.K.B., 83; 117, 191 Collins, A.J. See MacKay, N.J., 768 Collins, F.W. See Hutchinson, T.C., 606, 742; 763 Collins, J.D. See Segar, D.A., 563; 780 Collinson, R.I., 564, 643; 751 Cologer, C.P., 622; 751 Colwell, R.R., 112, 114; 117 See Austin, B., 558; 744 See Carney, J.F., 90, 95; 117 See Carty, C., 107; 117 See Hogan, M.A., 114; 117 See Oliver, J.D., 90; 123 See Quigley, M.M., 114; 124 See Schwarz, J.R., 111, 113; 125 Comes, R.A. See Bienati, N.L., 26; 33 Compton, K.G., 552, 589, 623; 751 Congleton, J.L., 427, 436; 464 Cornell, J.F. See Honjo, S., 82, 156, 168; 119, 192 Connor, P.M., 564, 730; 751 Connors, P.G. See Anderlini, V.C., 569; 745 Conover, R.J. See Clarke, G.L., 509; 529 Conrad, J.E., 261; 340 Contardi, V. See Capelli, R., 750 Cook, D.G., 205, 208, 216, 228, 232, 263, 264, 265, 287, 290, 296; 339 See Brinkhurst, R.O., 197, 320; 339 Cook, P. See Day, J.H., 406; 416 Cook, W.L. See Crow, S.A., 107; 119 Cooke, G.D. See Myers, D.N., 542; 772 Cooke, R.C. See Kepkay, P.E., 102; 122 Coombs, T.L., 549, 567, 611, 620; 751 See George, S.G., 757 Cooper, C. See McLachlan, A., 352; 417 Cooper, W.J., 88; 118 Copenhaver, E.D. See Braunstein, H.M., 539; 748 See Leland, H.V., 536, 601; 766 Copin-Montegut, C., 82; 118

Oceanography and marine biology Copin-Montegut, G. See Copin-Montegut, C., 82; 118 Copping, A.E., 79, 100; 118 Corcoran, E.F., 583; 751 See Alexander, J.E., 582, 635; 744 See Compton, K.G., 552, 589, 623; 751 Cordoba, A. See Speiss, F.N., 126 Cordoba, D. See Francheteau, J., 119, 757 Corliss, B.H., 142, 171; 191 Corliss, J.B., 115; 118 Cormier, M.J., 494, 497; 529 See Tsuji, F.I., 533 Corner, E.D.S., 564; 751 Cornet, D. See Delhaye, W., 545, 667, 734; 753 Corpe, W.A., 611, 616; 751 Corps of Engineers, San Francisco District, 752 Corre, M.C. See Bougis, P., 557, 566, 719; 748 Correll, D.L. See Faust, M.A., 86, 96; 119 See Friebele, E.S., 95; 119 Cosemans, G. See Kretzschmar, J.G., 586; 766 Cosgrove, E.G. See Oliver, B.G., 587; 774 Cossa, D., 571, 594; 752 Coston, L.C. See Hoss, D.E., 553; 762 Cothern, C.R. See Bergman, S.C., 593; 747 Coull, B.C., 336; 340 See Bell, S.S., 243, 281, 336; 338 Courtin, G.M. See Hogan, G.D., 548; 761 Courtois, L.A., 546; 752 Cowey, C.B., 458; 465 Cox, A. See Speiss, F.N., 126 Crabill, M.R. See Bryant, M.P., 105; 118 Craddock, J.E. See Backus, R.H., 528 Cragg, J.B. See Dash, M.C., 237; 263 Craig, H., 84, 178, 180; 118, 191 See Krishnaswami, S., 193 See Somayajulu, B.L.K., 179; 194 Craig, Jr, H.L., 623; 752 Craik, G.J.S., 450, 451; 465 Crane, J.M., 496; 529 See Cormier, M.J., 493; 529 Crane, K. See Corliss, J.B., 118 Cranston, R.E., 162; 192 See Emerson, S., 618; 754 Cranwell, P.A., 88, 89; 118 Crawford, C. See Grove, D.J., 458; 467 Crawford, C.C., 91, 93, 95; 119 Crecelius, E.A. See Gurtisen, J.M., 607; 759 See Young, J.S., 550; 789 Crenshaw, M.A. See Powell, E.N., 281; 347

815

Interlinking of physical

1290

Crepon, M., 57; 72 Crerar, D.A., 624; 752 See Means, J.L., 604; 769 Creutzberg, F., 446; 465 Crichton, M.O., 352, 355, 369, 374, 376; 416 Crill, P.A. See Soutar, A., 195 Crisp, M., 357, 392, 399; 416 Crisp, P.T. See Simoneit, B.R., 780 Critchlow, K.R., 425, 432, 443, 444, 449, 452, 456, 458, 459; 465 Cromer, J.L. See Shuman, M.S., 465; 780 Crothers, J.H. See Peden, J.D., 578; 775 Cross, F.A., 575, 580, 592, 595, 600, 619, 705, 712, 714; 752 See Huggett, R.J., 752; 762 Crow, S.A., 107; 119 Crowe, T.M. See Marsh, B., 422, 445, 461; 468 Cruz, A.A. de la See Odum, E.P., 87; 123 Cruz, L.L. See Montgomery, J.R., 771 Cuello, A.C., 31; 33 Cuhel, R.L. See Holm-Hansen, O.A., 1; 34 See Jannasch, H.W., 112; 121 Culkin, F. See Morris, R.J., 76; 123 Cullen, D.J., 281; 340 Cummins, K.W., 233, 241; 340 Cupka, D.M. See Anderson, Jr, W.D., 463 Curl, Jr, H. See Feely, R.A., 463, 600, 744; 755 Curl, H., 111; 119 See Nicholls, G.D., 574; 772 Currie, A.B. See Brown, A.C., 352, 404, 407; 416 Currie, R.I., 1, 12, 20; 33 Curtis, Jr, F.W., 33; 752 Curtis, L.A., 358; 416 Cuthbert, K.C., 352, 404, 408; 416 Cutshall, N., 594, 595, 639; 752 See Johnson, V., 594; 763 Dadone, A. See Frache, F., 757 Dagg, M.J., 78, 79; 119 D’Agostino, A., 565; 752 Dahl, I.O., 226, 228, 229, 231, 235, 236, 248, 271, 277, 280, 291, 321; 340 Dahlbäck, B.O., 108; 118 Dahlgren, U., 489, 606; 529 Dakin, W.J., 399, 400; 416 Dale, N.G., 237; 340 Dales, R.P., 320; 340 See Jørgensen, C.B., 312; 343 Daley, R.J., 86; 119 See Hobbie, J.E., 86; 119 See Overbeck, J., 99; 124

Oceanography and marine biology

817

Dallaire, G., 587; 752 Dallinger, R., 714; 752 See Wieser, W., 714; 787 Dalton, R.A. See Brown, V.M., 739; 749 D’Anglejan, B. See Subramanian, T., 514; 783 Daniel, M.J., 427, 433, 434, 436; 465 Danielsson, L.G., 560, 636, 637; 752 Darby, D.A. See Young, R.J., 613; 789 Darcy, G.H., 443, 457; 465 Darcy, K. See Reeve, M.R., 556; 777 Daris, F. See Majori, L., 566; 768 Darnell, R.M., 87; 119 Darracott, A. See Thornton, I., 576; 785 Dash, M.C., 237; 340 Dashora, M.S., 559, 720; 752 Da Silva, F.M. See Brown, A.C., 352, 388, 389, 392, 393, 410; 415 Datta Munshi, J.S. See Hughes, G.M., 436; 467 Daumas, R.A., 76; 119 Dauphin, J.P. See Beer, R.M., 160; 191 D’Auria, J.M. See Stump, I.G., 783 Dausend, K., 233, 303, 320; 340 Davenport, J., 427, 428, 437, 563, 717, 718, 733; 465, 752 See Crisp, M., 392; 416 See Manley, A.R., 544, 564, 717, 731, 732, 733, 738; 768 See Vahl, O., 457; 470 See Wirtz, P., 454; 471 Davey, E.W., 539, 554, 560, 592; 752 See Eisler, R., 754 David, C.N. See Clarke, G.L., 509; 529 David, P. See Benon, P., 747 Davies, A.G., 540, 612, 644, 716; 752 Davies, J.M. See Gamble, J.C., 743; 757 See Harrison, W.G., 540; 760 Davies, K.C. See Brown, A.C., 352, 404; 416 Davies, P.S. See Campbell, C.M., 431; 464 Davis, A.R. See Demayo, A., 600; 753 Davis, B.J., 423, 431; 465 Davis, E.M. See Guthrie, R.K., 556; 760 Davis, G. See Sieburth, J.McN., 126 Davis, G.K., 581, 589; 752 Davis, J.A., 752, 613, 615, 618; 752 See Leckie, J.O., 603; 766 Davis, J.C., 557; 752 Davis, J.S. See Gächter, R., 550; 757 Davis, M.H. See Romeril, M.G., 553, 589; 778 Davis, P.G. See Sieburth, J.McN., 126 Davis, P.H., 333; 340 See Spies, R.B., 333; 348

Interlinking of physical

1292

Davis, R.A. See Mathews, T.D., 659; 769 Davis, R.B., 281; 340 Davis, W.M. See White, D.C., 126 Davis, W.P., 592; 752 Dawes, E.A., 101; 119 Dawson, M.A. See Thurberg, F.P., 545; 785 Dawson, R., 91, 93, 96; 119 See Liebezeit, G., 76; 122 See Meyer-Reil, L.-A., 238; 345 Day, J.H., 406; 416 Dayal, R. See Duedall, I.W., 754 Dayton, P.K., 5; 33 Deacon, G.E.R., 1, 29; 33 Dean, J.M. See Cain, R.L., 422, 423, 449; 464 Deben, W.A. See Swartz, R.C., 628; 783 De Broyer, C., 31; 33 De Castro, G.L. See Montgomery, J.R., 770 Decleir, W., 537; 753 De Clerck, R., 659, 671, 684, 686, 692, 698, 699, 701, 704, 711, 712; 752 DeCoursey, P.J., 591; 753 Deetae, S. See Howe, M.R., 44; 73 Defant, A., 38, 57, 61; 72 De Ferret, C.A. See Heymer, A., 438, 439; 466 De Forest, A. See Pettis, R.W., 623; 775 Degens, E.T., 76, 84, 607; 119, 753 See Deuser, W.G., 84;118 See Ittekkot, V., 76; 121 See Siegal, A., 582; 780 De Goeij, J.J.M., 701; 752 De Groot, A.J., 595, 597, 629, 642; 752 See Salomons, W., 601; 778 Dehairs, F., 142; 192 Dehlinger, P., 192, 605, 617, 631, 636, 637, 663; 753 Dehorne, L., 209, 247; 340 De Jonge, V.N., 201; 340 De Kock, W.C. See De Wolf, P., 536; 752 Delach, M.N. See Horn, D.R., 624; 762 Delas, J. See Juste, C., 610; 764 Del Carratore, G. See Pace, F., 540; 774 De Leeuw, J.W. See Boon, J.J., 88; 117 De Leo, G., 439; 464 Delfino, J.J., 590; 753 Delhaye, W., 545, 667, 734; 753 Delisle, C.E., 582; 753 Dell, R.K., 1, 5, 31; 33 Delphy, M.J., 246, 247, 252, 254; 340 De Martini, E.E., 437, 440, 461; 464 See Marliave, J.B., 441; 468

Oceanography and marine biology Demayo, A., 600; 753 Demina, L.L., 594, 614, 617; 753 See Morozov, N.P., 753; 770 De Mora, S.J., 623; 752 Dempster, R.P., 622; 753 Demski, L.S., 503; 529 Den Dorren De Jong, L.E., 539, 557, 720; 753 Den Hartog, C., 245; 340 Dennon, Jr, T.J. See Groves, R.E., 622; 758 Denny, M., 391; 416 Denny, P. See Peter, R., 416; 775 Denton, E.J., 511, 514, 515, 518, 519, 522; 529 See Clarke, G.L., 523; 529 Denucé, J.M., 441; 465 Derenbach, J.B., 94, 97; 119 Derickson, W.K., 421, 422; 465 Desai, M.V.M., 601; 753 Descarries, L. See Anctil, M., 506; 528 Deschiens, R. See Floch, H., 563, 739; 756 Deshimaru, O., 538; 753 Dethier, M.N., 461; 465 Deuser, W.G., 84, 154, 157, 160; 119, 192 De Vega, V.R. See Ting, R.Y., 645; 784 Devol, A.H., 88; 118 See Christensen, J.P., 110; 117 De Vos, N., 269; 340 DeVries, A.L., 8, 11; 33 See Duman, J.G., 430; 464 De Wilde, P.A.W.J., 427; 464 De Wolf, P., 536, 621; 752 Dey, D.G. See Houlihan, D.F., 394; 417 Diachenko, G. See Sommer, S.E., 588; 781 Dias, J.K. See Anderson, Jr, W.D., 463 Diaz, R.J., 241, 245, 271, 281, 283, 291, 295, 321, 334, 335; 340 See Haven, D.S., 342 Diaz, R.K. See Anderson, JrW. D., 463 Diaz Garcia, V.M. See Speiss, F.N., 126 Dickman, M., 103; 119 Dickson, A. See Mantoura, R.F.C., 583; 769 Dickson, K.L. See Cairns, J., 561; 750 Diekmann, P. See Zeitschel, B., 154; 196 Dietz, A.S. See Sieburth, J. McN., 112; 125 See Yayanos, A.A., 111; 127 Dieulafait, M.L., 636; 753 Di Filippo, S. See Melluso, G., 769 Dinet, A. See Bodoy, A., 309; 338 Dinnel, P.A. See Stober, Q.J., 557; 782 DiSalvo, L.H. See Hirsch, N.D., 591; 761

819

Interlinking of physical Disteche, A. See Noel-Lambot, F., 574; 773 Dixon, P.N., 423; 465 See Milton, P., 423; 468 Dixon, P.S. See Christensen, E.R., 542, 587; 751 Djafari, D., 624, 643; 753 Djangmah, J.S., 549, 578, 684; 753 Doak, W., 440, 454, 480; 465, 530 Dodge, E.E., 579, 603; 753 Doi, K. See Kobayashi, N., 557; 765 Doig, III, M.T. See Martin, D.F., 543, 610; 769 Dokholyan, V.K. See Patin, S.A., 775 Dolar, S.G., 550; 753 Dolfing, J. See Zevenhuizen, L.P.T.M., 550; 789 Domey, R.G. See Davis, W.P., 592; 752 Dominik, J., 588; 753 Doneen, B. See Owens, A., 429; 469 Doneen, B.A., 429; 465 Dorband, W.R. See Ford, R.F., 756 Dörjes, J., 245, 284, 336; 340 Dotsu, U. See Kobayashi, T., 438, 448; 467 Dotsu, Y., 437, 440, 442; 465 See Shiogaki, M., 436, 441; 470 Dow, R.L., 544; 753 Downing, F.S. See Pyefinch, K.A., 753; 776 Doyle, L.J., 584; 754 Doyle, W.L., 428; 465 Dózsa-Farkas, K., 217; 340 Dradovskiy, S.G., 28; 33 Drawert, W., 313; 340 Drescher, H.E. See Harms, U., 570; 760 Drever, J.I., 600; 754 Driscoll, E.G., 281; 340 Drobrica, L. See Majtan, V., 101; 122 Druffel, E.M. See Williams, P.M., 84; 127 Dryssen, D., 127; 754 Duce, R.A., 585, 602, 632, 636; 754 See Piotrowicz, S.R., 585; 775 See Wallace, G.T., 108, 601, 605, 638; 126, 785 Duchart, P., 785; 754 Ducker, S, C. See Canterford, G.S., 582; 750 Dudley, W.C. See Dugolinsky, B.K., 624; 754 Duedall, I.W., 587, 636; 754 Duffrin, E. See Soutar, A., 195 Dugdale, R.C., 104, 119 See Barber, R.T., 604, 605; 747 See Packard, T.T., 104; 124 Dugger, W.M. See Gross, R.E., 540; 759 Dugolinsky, B.K., 624; 754

1294

Oceanography and marine biology Duinker, J.C., 593, 608, 613, 616, 635 637, 641; 754 Duke, T.W., 640; 754 See Cross, F.A., 580; 751 Duman, J.G., 430; 465 Dumnicka, E., 198; 340 Dunbar, R. See Dymond. J., 192 Dunker, S.S. See Burton, D.K., 550, 551, 559, 739; 750 Dunlap, J.C., 498; 530 Dunn, A. See Bebbington, G.N., 746 Dunne, J., 442, 458, 461; 465 Duplessy, J.-C. See Fontugne, M.R., 84; 119 Du Preez, H.H., 363, 365, 367; 416 Dürkop, H., 287; 340 Dusoge, K. See Kajak, Z., 201; 343 Dutton, J.W.R., 635, 640, 652, 655; 754 Duursma, E.K., 754, 602, 612; 754 See Aston, S.R., 594, 601; 744 Dybdahl, R. See Sale, P.F., 423; 470 Dybern, B.I., 201; 340 Dye, A.H., 352, 389, 394; 416 See McLachlan, A., 201; 345, 417 Dymond, J., 159; 191 Dymond, J. See Cobler, R., 83; 118, 191 Dymond, J. See Corliss, J.B., 118 Dzwillo, M., 197; 340 Eadie, B.J., 84; 119 Eaton, A., 594, 617, 632, 634; 754 Eaton, J.S., 162; 192 Ebeling, A.W., 433; 465 Eberhard, A., 476; 530 See Nealson, K.H., 476; 532 Eberhard, C. See Eberhard, A., 530 Ecker, R.M., 626, 644; 754 Edlund, A.-M., 459; 465 Edmond, J.M., 115; 118 See Bishop, J.K.B., 82, 84, 130, 142, 159 160, 163; 116, 191 See Boyle, F.A., 538, 628; 748 See Corliss, J.B., 117 Edwards, A.S., 510; 530 Edwards, C.A., 202; 340 Edwards, D.C. See Huebner, J.D., 365, 394; 417 Edwards, P. See Seelinger, U., 561, 610; 780 Edwards, R.W. See Learner, M.A., 325; 344 See Williams, N.V., 203, 217, 317; 349 Edwards, T.W. See Horn, M.H., 457; 467 Egami, F. See Hatanaka, H., 537; 761 Eger, W.G., 429, 432, 433; 465

821

Interlinking of physical

1296

Eggiman, D.W. See Betzer, P.R., 133; 191 Eglinton, G., 88, 89; 118 See Brooks, P.W., 117 See Cardoso, J., 88; 117 See Gaskell, S.J., 88; 119 See Morris, R.J., 76; 123 Ehlers, U. See Siebers, D., 240, 325, 544; 348, 780 Ehrlich, H.L., 624; 754 See Arcuri, E.J., 551, 624; 744 See Yang, S.H., 551, 558; 788 Ehrlich, K.F., 446; 465 Eichhorn, G.L., 611; 754 Eick, K., 204; 341 Eide, I., 554, 576; 754 See Myklestad, S., 561, 575; 771 Eimhjellen, K. See Jannasch, H.W., 110; 121 Eisenreich, S.J., 615; 754 Eisler, R., 539, 545, 562, 569, 571, 577, 589, 620, 659, 715, 716, 718, 732, 733, 734, 735; 754 See Joyner, T., 584; 764 See Mears, H.C., 699, 704, 710, 711; 769 Eisma, D., 141, 151; 192 Eittreim, S., 136; 191 Eittreim, S.L. See Biscaye, P.E., 134, 136; 191 See Ewing, M., 135; 192 See Jones, E.J.W., 135; 192 Elam, J.S., 462; 465 El-Barbary, I. See McDuffie, B., 592; 769 Elder, J.F. See Leland, H.V., 536, 600; 767 Elderfield, H., 767, 635, 638, 641; 755 See McArthur, J.M., 623; 769 Elias, R.W., 570, 579; 755 Eliason, A.H. See Smith, K.L., 126 Elizarov, A.A., 30; 34 Elkaïm, B. See Martoja, M., 585; 769 Elliott, H.A., 612, 614; 755 Elliott, J. See Frazer, J.Z., 757 Elliott, M. See McLusky, D.S., 321; 345 Ellis, D.V. See Coates, K., 324; 339 Ellis, J.P. See Sumraerhayes, C.P., 783 Elmgren, R., 201, 242, 243, 299, 300, 305, 309, 311, 336; 340 See, Ankar, S., 229, 230, 231, 233, 235, 271; 337 See Dybern, B.I., 201; 340 El-Sayed, S.Z., 1, 26; 33 See Holm-Hansen, O.A., 1; 34 Elson, P.F. See Sprague, J.B., 567; 782 El-Tawil, M.Y. See Miller, P.J., 442; 468 Eltink, A.T.G.W. See Creutzberg, F., 446; 465 Elvehjem, C.A., 536; 755

Oceanography and marine biology

823

Elzerman, A.W., 587, 606; 754 See Eisenreich, S.J., 615; 754 Emeh, C.O. See Menon, M.P., 595; 770 Emerson, R.R., 614; 755 See Harrison, F.L., 546, 589, 719, 743; 760 Emerson, S., 618, 639; 755 Emery, K.O., 142; 192 Emery, W.J. See Sievers, H.A., 29; 35 Engel, D.W., 603; 755 See Hoss, D.E., 552; 762 Engel, R. See Sick, L.V., 597; 780 Engle, R.L., 536; 755 Engvall, A.-G., 105; 119 Enomoto, N., 640, 661, 665, 670, 672, 681, 686, 695, 699; 755 Enright, J.T., 115; 119 Ensign, J.C., 102; 119 See Boylen, C.W., 101; 117 Environmental Protection Agency, Criteria and Standards Division, 742; 754 Eppley, R.W. See Harrison, W.G., 540; 756 Erasmus, T. See McLachlan, A., 260; 345, 417 Erez, J. See Honjo, S., 142, 171; 193 Erickson, S.J., 554, 718, 719, 723; 755 See Davey, E.W., 561; 752 Ericksson, E., 586; 755 Eriksson, L. See Wiederholm, T., 202, 203; 350 Erkenbrecher, C.W., 106; 119 Erlenkeu, H., 641; 755 Erman, D.C., 203; 340 Erman, N.A. See Erman, D.C., 203; 340 Ernst, W.H.O., 555, 576; 755 Erséus, C., 197, 198, 200, 204, 206, 208, 214, 228, 240, 254, 258, 263, 268, 274, 286, 290, 291, 294, 333, 337; 340 Erwin, J.A., 89; 119 Eshuis, E.J. See Zevenhuizen, L.P.T.M., 550; 789 Establier, R., 577, 578, 662, 669, 670, 680, 711; 755 Estabrook, R.W. See Werringloer, J., 539; 787 Estuarine Pollution Control and Assessment, 591; 755 Etienne, M. See Bougis, P., 557, 719; 748 Eustace, I.J., 697, 712; 756 See Ratkowsky, D.A., 776 See Thrower, S.J., 661; 784 Evans, D.H., 428; 465 See Wourms, J.P., 437, 441; 471 Evans, D.M. See Gordon, M.S., 465 Evans, D.W., 595; 756 Evans, III, E.C., 555; 756 Evans, E.C. See Hura, M., 539; 763 Evans, M.S. See Lewis, A.G., 578; 767

Interlinking of physical

1298

Evans, P.R., 244; 341 Ewing, J. See Ewing, M., 136; 192 See Jones, E.J.W., 135; 192 Ewing, M., 136; 192 See Eittreim, S., 135; 191 See Groot, J.J., 138; 192 See Jacobs, M.B., 133, 135, 137, 138; 192 See Jones, E.J.W., 135; 192 See Thorndike, E.M., 133, 135; 197 Eyles, J.C., 541; 756 Eyres, J.P., 575; 756 Fabris, G.J. See Harris, J.E., 556; 760 Fallis, B.W. See Bohn, A., 702, 712; 748 Fan, K.C., 162; 192 Fanger, H.U., 624; 756 Fanning, K.A. See Bender, M.L., 117 Farmanfarmaian, A. See Jannasch, H.W., 110; 121 Farrah, H., 592, 614, 618; 756 Farrell, M.P. See Winner, R.W., 557; 787 Farrington, J.W. See Goldberg, E.D., 758 See Jannasch, H.W., 82, 154; 121, 192 Fasching, C.P. See Piotrowicz, S.R., 586; 776 Fasching, J.L. See Duce, R.A., 754 Fassone, B. See Capelli, R., 750 Faubel, A., 239; 341 See Meyer-Reil, L.-A., 236, 238, 240; 345 Fauré-Frémiet, E., 260; 341 Faust, M.A., 86, 96; 119 See Friebele, E.S., 95; 119 Faust, S.L. See Fitzgerald, G.P., 568; 756 Fazio, S.D., 88; 119 Feder, H.M., 358; 416 See Naidu, A.S., 605; 771 Feely, H.W. See Li, Y.-H., 180; 194 See Sarmiento, J.L., 194 Feely, R.A., 82, 134, 135, 600, 744; 119, 192, 756 See Baker, E.T., 142; 190 See Landing, W.M., 190; 765 Feeney, R.E., 9; 34 Feig, Y.S. See Porter, K.G., 86; 124 Feller, R.J., 238; 341 Feltham, C.B. See ZoBell, C.E., 237; 350 Fenchel, T., 105, 235, 238, 257, 263, 276, 289, 291, 334; 119, 341 See Jørgensen, B.B., 106; 121 Fenn, W.O. See Marquis, R.E., 109; 123 Fenton, D. See Pamatmat, M.M., 113; 124 Ferguson, J. See Mooney, J.R., 619; 771

Oceanography and marine biology

825

Ferguson, J.F. See Bricker, O.P., 614; 749 Ferguson, R.L., 106; 119 See Palumbo, A.V., 106; 124 Fernandez, A. de S. See Alted, M.D., 422; 463 Fernandez, F.J. de S. See Alted, M.D., 422; 463 Ferrara, R. See Pace, F., 541; 774 Ferraris, J.D., 314; 341 Fewkes, R. See Sorem, R., 625; 781 Fiadeiro, M., 104; 119 Fiedler, J. See Blüm, V., 429; 464 Field, A.C. See Herbert, J.C., 714; 761 Fielden, J.M. See Leland, H.V., 571; 767 Fielding, A.H., 542; 756 Filimonov, V.S., 511; 530 Filip, D.S., 580, 721; 756 Findley, L.T. See Thomson, D.A., 422; 470 Fink, Jr, L.K., 470, 643, 661, 665, 677; 755 Finkel, A.J., 566, 740; 756 See Allee, W.C., 566; 744 Finney, C. See D’Agostino, A., 565; 752 Finogenova, N.P., 291; 341 Fischer, K. See Dymond, J., 192 Fischer, S. See Gordon, M.S., 433; 466 Fisher, D.E. See Bonatti, E., 614; 748 Fisher, J.A., 280; 341 Fisher, J.B. See McCall, P.L., 280, 281, 322; 345 Fisher, N.S., 555; 756 See Bankston, D.C., 747 Fisheries Department Fishery Data Centre. Food and Agricultural Organization, 689; 756 Fisk, M.B. See Frazer, J.Z., 757 Fitch, J.E., 420; 465 Fitzgerald, G.P., 568; 756 Fitzgerald, J.M., 484; 530 Fitzgerald, M.G. See Summerhayes, C.P., 783 Fitzgerald, W.F. See Dehlinger, P., 754 Fives, J.M., 440, 442, 457,404, 461; 465 Fleer, A. See Spencer, D.W., 83; 126, 195, 781 Fleer, A.P. See Brewer, P.G., 191 Flegal, A.L. See Martin, J.H., 649; 769 Flegal, A.R., 549, 659, 675; 756 Fleming, T.P., 325; 341 See Richards, K.S., 291; 347 Fletcher, G.L., 538, 574, 575, 703, 713; 756 Fletcher, I.S. See Duce, R.A., 754 Fletcher, K. See Grieve, D., 595, 598; 759 Fletcher, W.K. See Grieve, D., 595, 614; 759 Floch, H., 563, 739; 756 Floodgate, G.D. See Litchfield, G.D., 76, 106; 122

Interlinking of physical

1300

See Nedwell, D.B., 106; 123 Florence, T.M., 616; 756 See Batley, G.E., 601, 608, 614, 618; 747 Foberg, M., 341 Fogg, G.E., 20; 34 Folger, D.W., 139; 192 Folkard, A.P. See Dutton, J.W.R., 754 Follett, W.I., 461; 465 Fomina, L.S. See Demina, L.L., 594, 614, 617; 753 Fonds, M., 422, 424, 428, 430; 466 Fonselius, S.H., 636; 756 Fontugne, M.R., 84; 119 Forbes, M.A. See Demayo, A., 600; 753 Ford, N. See Le Ray, W., 203; 344 Ford, R.F., 636, 639; 756 Fordyce, R.E., 29; 34 Foret, J.P. See Bellan, G., 328; 338 Forster, G.R. See Pingree, R.D., 543; 776 Forster, W.O., 637; 756 Förstner, U., 756, 617, 621, 629, 640, 643 See Dominik, J., 587; 753 See Müller, G., 613, 642; 771 See Patchineelam, S.R., 642; 775 See Stoffers, P., 614; 782 Foster, M.A., 427; 466 Foster, P., 561, 577, 578, 582, 614, 615, 618, 635, 637, 643; 757 See Walker, G., 546, 548, 550, 584, 681; 785 Foster, P.L., 548; 757 Foster, R.B., 575, 577; 757 Foster, R.S. See Neff, J.W., 582; 773 Foster, T.D., 26, 29; 33 See Carmack, E.C., 29; 33 Fouda, M.M., 443; 466 Fowler, S.W., 81, 166, 564, 566, 616, 644, 647, 659, 669; 119, 192, 757 See Beasley, T.M., 190 See Cherry, R.W., 185; 191 See Komar, P.D., 163, 165; 193 See Small, L.F., 165, 580; 194, 780 Fox, H.M., 302, 320; 341 Fox, I. See Frick, L.P., 563; 757 Fox, L.S., 422; 466 Fox, P.J. See Francheteau, J., 119, 757 Foye, W.O., 622; 757 Frache, F., 635; 757 Francescon, A. See Barbaro, A., 746 Franceshini, G.A. See Holm-Hansen, O.A., 1; 34 Francheteau, J., 115, 625; 119, 757 See Speiss, F.N., 125

Oceanography and marine biology Frankenne, F. See Noel-Lambot, F., 574; 773 Franks, R., 636, 639; 757 Franks, R.P. See Bruland, K.W., 159; 191 Fraser, J., 515; 530 Fraser, J.H., 258, 322; 341 Frassetto, R., 57; 72 Frazer, J.Z., 600, 624; 757 Frazier, J.M., 539; 757 Fredericks, J. See Brewer, P.G., 191 Freiberger, A. See Cologer, C.P., 622; 751 French, F.W. See Sieburth, J. McN., 126 Fretter, V., 369; 416 Friberg, E. See Mague, T.H., 78; 122 Frick, L.P., 563; 757 Fricke, A.H., 202; 341 Fridman, D., 480; 530 Friebele, E.S., 96; 120 Frieden, E., 537; 757 Friedland, A.C. See Lev-er, J., 589; 767 Froelich, P.N. See Bender, M.L., 117 Frolova, E.N., 247, 248; 341 Fry, F.E.J., 717; 757 Fryer, J.L. See Hetrick, F.M., 546; 761 Fryer, J.N., 429; 466 Fuhrman, F.A. See Elam, J.S., 462; 465 Fuhrman, J.A., 98; 120 See Hollibaugh, J.T., 97; 119 Fujiya, M., 719, 732, 734; 757 Fukai, R., 636; 757 Fukuda, M. See Jerlov, N.G., 518; 531 Funk, W.H. See Bartlett, L., 619; 746 Fürstenberg, J.P. See McLachlan, A., 260; 345 Gaboury, G.A. See Eisler, R., 539; 755 Gächter, R., 541, 550, 606, 620; 757 Gadre, G.T. See Kappana, A.N., 764 Gagner, C. See Bender, M.L., 538, 636; 747 Gallagher, E.D. See Feller, R.J., 341 Gallagher, J.L.,98; 120 See Wolf, P.L., 576, 578; 788 Gallardo, V.A., 116; 120 See Morita, R.Y., 116; 123 Galli-Galardo, S.M. See Marusic, E.T., 468 Galloway, J.N., 587, 597, 631, 643; 757 Galloway, W. See Widdows, J., 556; 787 Galtsoff, P.S., 787, 584, 636; 757 Gambell, R., 8; 34 Gamberoni, L. See Lacombe, H., 57; 73

827

Interlinking of physical

1302

Gamble, E. See Goldberg, E.D., 588; 759 Gamble, J.C., 743; 757 See Reeve, M.R., 544; 777 Gambrell, R.P. See Khalid, R.A., 616; 765 See Patrick, Jr, W.H., 601; 775 Ganapati, P.N., 297; 341 See Subba Rao, B.V., 219, 226, 297, 313, 323; 348 Ganguly, A.K. See Desai, M.V.M., 601; 753 Ganning, B., 426; 466 Garcia, R.J. See Lowman, F.G., 768 Gardener, W.S., 333; 341 Gardner, D. See Batley, G.E., 594, 614; 746 Gardner, G.R., 545, 569; 757 See Eisler, R., 544, 568, 619, 620, 718, 735; 754 Gardner, L.R., 579, 588, 642, 658; 757 Gardner, W. See Dymond, J., 192 Gardner, W.D., 131, 145, 154, 160, 163; 192 See Reynolds, C.S., 154; 194 See Rowe, G.T., 154, 157; 194 Gardner, W.S. See Hanson, R.B., 106; 120 Garey, J.F., 623; 757 Gariépy, P., 506; 530 Garner, H.R. See Allee, W.C., 566; 744 See Finkel, A.J., 566; 755 Garland, T.R. See Schmidt, R.L., 579; 779 Garreton, M. See Marusic, E.T., 468 Garrett, W.B., 107; 120 Gartner, S. See Rezak, R., 777 Gartside, D.W., 548; 758 Garvine, R.W. See Dehlinger, P., 753 Gascard, J.C. See Lacombe, H., 38, 60; 73 Gaskell, S.J., 88; 120 See Brooks, P.W., 117 Gassmann, G. See Gunkel, W., 332; 342 Gat, Y. See Lev-er, J., 589; 767 Gaudette, H.E. See Armstrong, P.B., 588; 745 See Lyons, W.B., 625, 641; 767 Gauld, D.T., 352; 416 Gauthier, M. See Aubert, M., 745 Gauthier, M.J., 542, 620; 758 Gedye, A. See Brown, A.C., 397, 408; 416 Geladi, P. See Decleir, W., 537; 753 Geldiay, R., 563, 672; 758 Gelpi, E. See Tornabene, T.G., 88; 127 Genetelli, E.J. See Gould, M.S., 609, 616; 759 Gentile, J.H. See Sosnowski, S.L., 552, 556, 565, 727, 744; 781 Gentry, J. See Fan, K.C., 162; 192 George, S.G., 546, 644, 669, 670; 758

Oceanography and marine biology

829

See Coombs, T.L., 548; 751 Georgi, D.T. See Gordon, A.L., 29; 34 GEOSECS, 130; 192 Gerchakov, S.M., 590, 622; 758 Gerday, C. See Noel-Lambot, F., 574; 773 Gerking, S.D., 450, 451; 466 Gerlach, S.A., 237, 241, 271, 282, 336, 642; 341, 758 Germond, D.J. See Sosnowski, S.L., 552, 744; 781 Gersch, M. See Richter, K., 313; 347 Ghiorse, W.C., 625; 758 Ghosh, M.M. See Olofsson, J.A., 600; 774 Ghuman, G.S. See Menon, M.P., 595; 770 Gibbons, J.S., 399; 416 Gibbs, P.E., 536; 758 Gibbs, R.J., 130, 592; 192, 758 Gibson, C.I., 564, 718, 719, 731; 758 See Schmidt, R.L., 758, 613, 642; 778 Gibson, R. See Jennings, J.B., 244; 343 Gibson, R.N., 420–71; 420, 421, 422, 423, 425, 427, 431, 440, 442, 443, 444, 448, 449, 451, 452, 453, 454, 455, 456, 457, 459, 460, 462; 466 See Kislaliogu, M., 422, 456, 457, 460; 466 Gibson, V.R., 566; 758 See Reeve, M.R., 776 Giere, O., 197–350; 202, 203, 204, 217, 219, 223, 225, 226, 228, 229, 230, 231, 234, 236, 237, 238, 239, 240, 241, 248, 252, 254, 257, 259, 260, 263, 267, 268, 269, 274, 276, 279, 283, 284, 287, 288, 289, 291, 297, 299, 300, 301, 303, 306, 309, 310, 323, 326, 327, 333, 334, 335, 336, Fig. 10 (Giere & Pfannkuche) facing p. 252, Fig. 37 (Giere & Pfannkuche) facing p. 331; 341 Giese, G.S. See Steele, J.H., 201, 236; 348 Gieskes, J.M., 40; 72 Gieskes, W.W.C. See Eisma, D., 141, 151; 192 Giesy, Jr, J.P., 192; 758 Gilchrist, J.D.F., 362, 381; 416 Giles, J. See Timoney, J.F., 548; 785 Gillian, F.T. See Volkman, J.K., 127 Gillain, G. See Billen, G., 91; 117 Gilles, R., 537; 758 Gillespie, P.A., 91, 554, 559, 605; 120, 758 See Sunda, W.G., 553, 558; 783 Gilmartin, M., 697, 707; 758 Gilpin-Brown, J.B. See Denton, E.J., 511, 519; 529 Gingi, D.T. See Gordon, A.L., 1, 28; 34 Ginsler, P.J. See Kleerekoper, H., 567; 765 Girvin, D.C., 592, 594; 758 See Anderlini, V.C., 744 Gjessing, E.T., 602, 605, 609; 758 Glasby, G.P., 623, 624; 758 Glaser, O., 538, 566, 740; 758 Glass, G.E. See Magnuson, V.R., 768

Interlinking of physical

1304

Glassen, R.C. See Pilotte, J.O., 776 Glazunov, V.A. See Lisitzin, A.P., 133; 194 Glover, J.W., 699, 702; 758 Gluth, G. See Gnaiger, E., 274; 342 Gnaiger, E., 274; 342 See Wieser, W., 274; 349 Gnassia-Barelli, M., 560, 606; 758 Gocke, K., 91, 95; 119 See Dawson, R., 90, 92, 95; 118 Goddard, J. See Bacon, M.P., 190 Godoy, O.T., 578; 758 Goering, J.J., 104, 107, 540; 119, 758 See Dugdale, R.C., 118 See Packard, T.T., 104; 124 Goerke, H., 243; 342 Goldberg, E.D., 342, 556, 570, 581, 587, 600, 617, 619, 628, 630, 632, 640, 642, 643, 690, 693, 712; 192, 758 See Bertine, K.K., 586, 587; 747 See Bruland, K.W., 191 See Hodge, V., 762 See Wood, J.M., 638; 788 Goldberg, R.D. See Gordon, A.L., 16; 34 Goldhaber, M.B., 105; 120 Goldman, C.R., 538; 759 See Kimmel, B.L., 154; 192 See Paerl, H.W., 92; 124 Goldschmid, A., 457; 466 Gollub, A.R., 455; 466 Golombick, T., 352, 404, 407; 416 Goodell, H.G. See Pilkey, O.H., 574; 775 Goodgal, S. See Carritt, D.E., 613, 617; 750 Goodman, C., 555; 759 Gordeyev, E.I., 139; 192 Gordeyev, V.V. See Demina, L.L., 594, 617; 753 Gordiyenko, S.A. See Lasik, Y.A., 611; 766 Gordon, A.L., 1, 16, 26, 28; 34 See Eittreim, S., 191 Gordon, D.C., 82, 84; 120 Gordon, H.R., 134; 192 Gordon, L.I. See Corliss, J.B., 118 See Williams, P.M., 84; 127 Gordon, M.S., 433, 434, 436, 456; 466 Goreau, T.J., 171; 192 Goring, C.A.I., 103; 120 Gorkin, I.N. See Voronina, E.A., 547, 569, 740; 786 Goto, T. See Inoue, S., 497; 531 See Tsuji, F.I., 533 Goulart, E. See Brockamp, O., 624; 749

Oceanography and marine biology

831

Gould, M.S., 610, 616; 759 Goulder, R., 87, 554, 559, 718; 120, 759 Goupil, D.W. See Baier, R.E., 108; 117 Goyer, R.A. See Cherian, M.G., 550; 751 Grachev, Y.P. See Zhirova, V.V., 559; 789 Graham, A. See Fretter, V., 369; 416 Graham, B.W. See Wagemann, R., 592; 786 Graham, D.L., 666, 669, 675, 680; 759 Graham, J.B., 431, 433, 435, 456, 462; 466 Graham, K.J. See Leumer, G., 590; 767 Graham, P.H., 484; 530 Grainger, E.H., 28; 34 Grant, P.T. See George, S.G., 758 Grant, V. See Carpenter, J.H., 585; 750 Grassle, J., 273, 312; 342 Grassmann, E. See Kirchgessner, M., 619; 765 Grave, C., 538, 566; 759 Gray, J.S., 230, 233, 237, 244, 304, 321, 336; 342 See Hulings, N.C., 259, 288, 336; 342 See Uhlig, G., 203; 349 Gray, R.W. See Williams, P.J. Le B., 91; 127 Gray, T.R.G., 94, 112; 119 Green, J.M., 423, 424, 427, 432, 443, 444, 449, 450, 451, 453, 455; 466 Green, K. See Corliss, J.B., 118 Greenberg, E.P. See Ruby, E.G., 477, 491; 533 Greene, C.S., 587; 759 Greene, C.W., 499, 508; 530 Greene, H.H. See Greene, C.W., 499, 508; 530 Greene, J.C. See Miller, W.E., 555; 771 Greenwood, P.H., 474; 530 Greenwood, P.J. See Brown, A.C., 408; 416 Gregory, R.B., 434; 466 Greig, R.A., 567, 575, 576, 580, 643, 644, 646, 700, 702, 706, 712; 759 See Wenzloff, D.R., 570; 785 Greter, F.L. See Buffle, J., 609; 749 Grice, G.D. See Gibson, V.R., 566; 758 See Reeve, M.R., 776 Grieshaber, M. See Pörtner, H.-O., 320; 347 Grieve, D., 595, 598, 614; 759 Griffin, J. See Goldberg, E.D., 759 Griffin, J.J. See Goldberg, E.D., 588; 759 Griffith, E., 351; 416 Griffith, R.W. See Marusic, E.T., 468 Griffiths, A. See Chester, R., 751 Griffiths, D., 31; 34 Griggs, G.B., 587, 643; 759 Grill, E.V. See Thomas, D.J., 595, 637; 784 Grimanis, A.P. See Zafiropoulos, D., 647; 789

Interlinking of physical

1306

Grizzard, T.J. See Randall, C.W., 588; 776 Grondey, G. See Fanger, H.U., 624; 756 Grontved, J. See Bertelsen, E., 499, 511; 529 Groot, J.J., 138; 192 Groover, R.E., 590, 622; 759 Gross, M.G., 558, 587, 742; 759 See Ali, S.A., 626; 744 Gross, R.E., 540; 759 Grossman, G.D., 423, 442, 452, 453, 457, 458, 460, 463; 467 Group of Experts on the Scientific Aspects of Marine Pollution, 623; 758 Grove, D.J., 458; 467 See Djangmah, J., 548, 683; 753 Groves, R.E., 622; 759 Gruchy, C.G. See Anctil, M., 509; 528 Gruger, Jr, E.H. See Karrick, N.L., 539, 719, 739; 764 Grundmanis, V., 103, 105; 120 See Murray, J.W., 113; 123 Gruzov, E.N., 1, 5, 28, 31; 34 Guard, C.L. See Murrish, D.E., 9; 35 Guarino, C.F., 588; 759 Gudat, J.C. See Cardeilhac, P.T., 750 Guerin-Ancey, O.J. See Berland, B.R., 541; 747 Guerrero, J. See Francheteau, J., 119, 757 See Speiss, F.N., 125 Guest, W.C. See Roegge, M.A., 622; 778 Guillard, R.R.L. See Anderson, M.A., 603; 745 See Bankston, D.C., 747 See Degens, E.T., 84; 118 See Morel, N.M.L., 537; 770 See Murphy, L.S., 548; 771 See Sunda, W., 551, 578, 582, 601, 602, 723; 782 Gulland, J.A., 6; 34 Gunam, N.S. See Milanovich, F.P., 658; 770 Gundersen, K., 106; 120 Gundlach, H. See Schnier, C., 625; 779 Gunkel, W., 332; 342 Gunn, V.P. See De Goeij, J.J.M., 753 Gunner, H.B. See Berger, P.S., 612; 747 See Coler, R.A., 238; 339 Gunther, R.L. See Fryer, J.N., 429; 466 Gupta, A.B., 543; 759 Gupta, S.K., 614, 641, 644; 759 Gupta, R.S. See Dashora, M.S., 559, 720; 752 Gurtisen, J.M., 607; 759 See Young, J.S., 789 Guthrie, R.K., 556, 576, 661, 663, 671, 682; 760 See Cherry, D.S., 576; 750 See Rodgers, J.H., 561; 777

Oceanography and marine biology

833

See Singleton, F.L., 548, 558; 780 Gutierrez, M. See Establier, R., 669, 712; 755 Haedrich, R.L. See Backus, R.H., 528 Haerdi, W. See Buffle, J., 609; 749 Hagen, G., 199, 258, 264, 269, 271, 275, 280; 342 Hagmeier, E., 139; 192 Hagström, A. See Larsson, U., 94, 97; 122 Haight, R.D. See Morita, R.Y., 109; 123 Haines, J.L., 267; 342 Haka, P., 233; 342 Hakala, I., 201; 342 Håkansson, N. See Kjelleberg, S., 108, 109; 122 Halbach, P., 624; 760 Halcrow, W., 635, 643, 647, 683, 685, 690, 691, 692; 760 Halkyard, J.E., 625; 760 Hall, G.L. See Bouma, A.M., 627; 748 Hall, K.H. See Benedict, A.H., 592; 747 Hall, K.J. See Koch, F.A., 588; 765 Hall, R.A., 644, 661, 681, 689; 760 Hallberg, R., 607; 760 Hallberg, R.O., 760, 641, 642; 760 See Engvall, A.-G., 106; 118 See Mooney, J.R., 618; 770 Halverson, R. See Case, J.F., 529 Hamilton, R.D., 88, 91; 120 Hamilton-Taylor, J., 606; 760 Hamlin, J.M. See Jackim, E., 568; 763 Han, S.L.M. See Hung, T.C., 615; 763 Handler, P. See White, A., 537; 787 Haneda, Y., 473, 479, 481, 483, 484, 485, 486, 487, 491, 492, 493, 494, 496; 530 See Abe, T., 486; 529 See Johnson, F.H., 492, 493; 531 See Shimomura, O., 497; 532 See Sie, E.H.-C., 493; 532 See Tsuji, F.I., 493, 494, 497; 533 See Yoshiba, S., 483, 485; 533 Hanke, W. See Muller, R., 429; 469 Hanks, K.S., 731; 760 Hanley, A. See Brewer, P.G., 191 Hanna, B.A., 561; 760 Hannan, P.J., 554, 722, 739; 760 Hansen, K., 489; 531 Hanson, C.H., 623; 760 Hanson, E.H. See Schnitzer, M., 604; 779 Hanson, G.M. See Armstrong, P.B., 588; 745 Hanson, H.C., 576; 760 Hanson, P.J. See Sunda, W.B., 603, 604; 783

Interlinking of physical

1308

Hanson, R.B., 106; 120 See Briggs, K.B., 238; 338 Hantschmann, D. See Krey, J., 193 Hanus, F.J. See Baross, J.A., 109; 117 Hanya, T. See Ishiwatari, R., 88, 90; 121 Harder, H. See Brockamp, O., 624; 749 Harding, G.C.H., 82; 120 Harding, J.P.C., 551; 760 Hardisty, M.W., 713; 760 Harden, M.J. See Albright, L.J., 109; 116 Hardy, L.H. See Cross, F.A., 575; 752 Hargrave, B.T., 154, 158, 160; 192 Harman, W.J. See Loden, M.S., 199; 345 Harms, U., 570; 760 Harnisch, O., 276, 302, 320; 342 Harrell, R.C. See Olson, K.R., 562, 718, 733; 774 Harriman, D.M. See Dow, R.L., 544; 753 Harrington, A. See Morin, J.G., 532 Harris, A.B. See Fink, Jr, L.K., 532; 755 Harris, E.J. See Zaba, B.N., 573; 789 Harris, J.E., 82, 142, 150, 556, 563, 579, 656 657, 668, 689; 119, 192, 760 See Bassin, N.J., 133; 190 See Feely, R.A., 82; 118 Harris, W.H. See Ramondetta, P.J., 642; 776 Harrison, B.J. See Harrison, P.G., 233 235, 241; 342 Harrison, F. See Orlob, G.T., 604; 774 Harrison, F.L., 547, 553, 556, 560, 576, 580 581, 589, 590, 602, 659, 718, 720, 732 739; 760 See Emerson, R.R., 613; 754 See Knezovich, J.P., 550, 731; 765 See Rice, Jr, D.W., 548, 567, 734, 735777 Harrison, P.G., 233, 235, 241; 342 Harrison, W.G., 540; 760 Harriss, D.K. See Magnuson, V.R., 770 Harriss, R.C., 581; 761 See Horvath, G.J., 761; 762 Harry, H.W. See Yager, C.M., 581788 Hart, R.C. See McCapra, F., 497; 532 Hart, T.J., 20; 34 Hartwig, E., 247, 248, 254; 342 See Pfannkuche, O., 245, 271, 296; 345 Hartwig, E.O. See Carlucci, A.F., 103; 118 Harvey, B.R. See Hetherington, J.R., 643; 761 Harvey, E.N., 473, 483, 492, 499, 508, 511; 531 Harvey, G. See Goldberg, E.D., 758 Harvey, G.R. See Stuermer, D.H., 609 610; 783 Harvey, G.W., 107; 120 Harvey, H.W., 716; 761 Hashimoto, J. See Hori, K., 467

Oceanography and marine biology

835

Haskins, C.P. See Hutner, S.H., 538; 763 Hasle, G.R., 20; 34 Hass, G., 242; 342 Hassall, K.A. See McBrien, D.C.H., 539; 769 Hasse, L. See Slinn, W.G.N., 781 Hassett, J.M. See Jennett, J.C., 560; 764 Hastings, J.W., 486, 490, 523; 530 See Dunlap, J.C., 497; 529 See Morin, J.G., 531 See Nealson, K.H., 476, 490; 532 See Reichelt, J.L., 486; 532 See Ruby, E.G., 476, 490; 532 See Ulitzur, S., 476; 533 Hata, Y., 558, 721; 761 Hatanaka, H., 537; 761 Hathaway, J. See Meade, R.H., 123 Hattori, A. See Goering, J.J., 541; 758 See Koike, I., 103, 104, 106; 122 See Saino, T., 84; 124 See Seki, H., 110; 125 Hattori, J. See Sasaki, T., 424, 443, 450, 460; 470 Haug, A., 554, 583, 648, 652; 761 Haugsness, J.A. See Smith, K.L., 114; 125 Hauschildt, D., 217, 218, 236, 284, 326, 329; 342 See Giere, O., 203, 204, 217, 223, 230, 231, 236, 238, 268, 325, 326, 333, Fig. 37 (Giere & Pfannkuche) facing p. 330; 341 Hautekiet, W. See Heip, C., 201; 342 Haven, D.S., 229; 342 Hawkins, J. See Speiss, F.N., 126 Hayashi, M., 422, 423; 467 Hayes, W.B., 687; 761 Haymon, R. See Speiss, F.N., 126 Hazel, C.R., 569, 736; 761 Head, R.N. See Pingree, R.D., 543; 776 Heald, E.J., 239; 342 Healey, M.C., 442, 457, 459, 460; 467 Healy, M.L. See Lieberman, S.H., 607, 613; 767 See Packard, T.T., 106, 110; 124 See Sugai, S.F., 635; 782 See Zirino, A., 601, 602; 789 Heath, A.G. See Cairns, J., 551; 750 Heath, G.R. See Bender, M.L., 117 See Hinga, K.R., 192 Heath, W.A. See Bawden, C.A., 592; 746 See Parsons, T.R., 612; 774 Hedges, J.I. See Rau, G.H., 115; 125 Hedgpeth, J.W., 1, 29; 34 Heesen, T.C. See Young, D.R., 590; 789

Interlinking of physical

1310

Heezen, B.C., 63; 73 See Folger, D.W., 138; 192 Heffter, J.L. See Duce, R.A., 754 Heggie, D.T., 595, 597, 636, 640, 641; 761 See Reeburgh, W.S., 106; 124 Heidenreich, E., 248; 342 Heip, C., 201, 210, 227, 235, 269; 342 Heitzer, R.D., 105; 120 Hekinian, R. See Francheteau, J., 119, 757 Helene, C. See Bere, A., 611; 747 Hellebust, J.A. See Degens, E.T., 84; 119 Helmy, M.M., 544; 761 Helz, G.R., 617, 643; 761 See Hill, J.M., 658; 762 Hemmingsen, A.M., 394; 416 Hempel, G. See Hempel, I., 29; 34 Hempel, I., 29; 34 Henderson, J.S. See Minear, R.A., 771 Hendricks, J.W. See Ecker, R.M., 626, 644; 754 Hendrie, M.S., 475; 531 Henigman, J.F. See Albright, L.J., 109; 117 Henningsson, B. See Norman, E., 622; 773 Henriksen, K., 103; 120 Henry, J.P., 477; 531 Hepworth, A. See Elderfield, H., 755 Herald, E.S., 435, 460; 467 Herald, O.W. See Herald, E.S., 435, 460; 467 Herbert, J.G., 714; 761 Herbich, J.B., 591, 627; 761 Herdson, D.M. See Evans, P.R., 244; 341 Heritage, J., 587; 761 Herlant-Meewis, H., 209; 342 Hermansson, M. See Dahlbäck, B.O., 108; 119 Herring, P.J., 473–533; 473, 480, 483, 485, 487, 488, 489, 497, 511, 514, 515, 519, 522, 523; 530, 531 See Denton, E.J., 523; 529 See Edwards, A.S., 510; 529 See Hansen, K., 488, 489; 530 Hershelman, G.P., 587; 761 Herzen, R.P.von See Corliss, J.B., 118 Heslinga, G.A., 717; 761 Hess, S. See Craig, Jr, H.L., 752 Hesser, R.R. See Enright, J.T., 115; 119 Hessler, R. See Speiss, F.N., 126 Hestand, R.S., 542; 761 Hester, B.S. See Davis, W.P., 592; 755 Hesthagen, I.H., 425, 430, 443, 455; 466 See Berge, J.A., 243, 460; 338, 463

Oceanography and marine biology See Gibson, R.N., 454, 455; 465 Hetherington, J.A., 643; 761 Hetrick, F.M., 546; 761 Heuss, E., 608; 761 Hevert, F. See Schipp, R., 573, 674; 779 Heydemann, A. See Brockamp, O., 623; 749 Heymer, A., 439, 451; 467 See Zander, C.D., 443, 456; 471 Heyraud, M., 185; 193 See Beasley, T.M., 190 See Cherry, R.W., 185; 191 Hiby, A.R. See Stebbing, R.D., 548, 726, 739; 782 Hicks, B.B. See Slinn, W.G.N., 781 Hidaka, I., 569; 761 Higgins, H.W. See Smith, D.F., 126 Higgo, J.J.W. See Beasley, T.M., 191 Higgs, T.W., 587; 761 Hill, J.M., 658, 761 See Helz, G.R., 617; 762 Hill, S.R. See Nygaard, D.D., 600; 773 Hiltner, R.S., 636; 761 Hiltunen, J.K. See Cook, D.G., 205, 296; 340 Hinga, K.R., 156; 193 See Sieburth, J.McN., 125 Hinton, P.J. See Black, G.A.P., 558; 747 Hiraizumi, Y., 642; 761 Hirayama, K., 543; 761 Hirobe, T. See Isibasi, M., 763 Hirobe, Y. See Nozaki, Y., 183; 194 Hirsch, N.D., 591; 761 Hissong, D.E. See Pringle, B.H., 689; 776 Hjort, J. See Murray, J., 515; 532 Hobbie, J.E., 86, 88, 91, 106, 240; 119, 343 See Crawford, C.C., 90; 118 See Daley, R.J., 86; 118 See Wright, R.T., 90, 91, 97; 127 Hobson, E.S., 421; 467 Hobson, L.A., 82; 120 Hochman, H., 590; 761 Hodge, V., 634; 761 See Goldberg, E.D., 758 Hodgkiss, W., 85; 120 See Hendrie, M.S., 474; 530 Hodgson, A.T. See Girvan, D.C., 592; 758 Hodson, R. See Thomas, W.H., 784 Hodson, R.E., 88; 120 See Vaccaro, R.F., 558; 785 Hodson, P.V., 539; 761

837

Interlinking of physical

1312

Hoehn, R.C. See Randall, C.W., 588; 776 Hoek, G.J.V. See Boon, J.J., 89; 117 Hoff, J., 594; 761 Hoffman, E.J. See Duce, R.A., 754 Hoffman, G.L. See Duce, R.A., 585; 754 See Piotrowicz, S.R., 775 See Rheinberger, R., 584; 777 See Wallace, Jr, G.T., 586; 785 Hoffman, K.H., 320; 343 Hogan, A.W. See Slinn, W.G.N., 781 Hogan, G.D., 548; 761 Hogan, M.A., 114; 120 Hoigne, J., 539; 762 Holdgate, M.W., 1, 2, 5; 34 Holl, A. See Schulte, E., 439; 470 Holland, J.F. See Cline, J.T., 602, 614; 751 Hollibaugh, J.T., 96; 120 See King, K.R., 101; 121 See Thomas, W.H., 783 Holligan, P.M. See Pingree, R.D., 543; 776 Hollister, C.D. See Gardner, W.D., 192 Hollod, G.J. See McDuffie, B., 592; 770 Holloway, P.J. See Cardoso, J., 89; 118 Holmes, C.W. See McLerran, C.J., 624; 770 Holm-Hansen, O., 88; 119 See Berman, T., 76; 116 See Devol, A.H., 87; 118 See Hamilton, R.D., 87; 119 See Hobbie, J.E., 119 See Hodson, R.E., 87; 119 See Karl, D.M., 87, 106; 121 See Thomas, W.H., 540; 783 See Williams, P.J. Le B., 90, 94; 127 Holm-Hansen, O.A., 1, 21, 26; 34 Holm-Jensen, I.B., 564, 718, 730; 762 Holmquist, C., 205; 343 Holopainen, I.J., 201; 342 See Haka, P., 342 Holton, R. See Cutshall, N., 594; 752 Holtzclaw, K.M. See Sposito, G., 610; 782 Honjo, S., 82, 83, 142, 155, 156, 159, 167, 168, 171; 121, 192 See Corliss, B.H., 142, 170; 191 See Emery, K.O., 142; 191 See Goreau, T.J., 171; 192 See Spencer, D.W., 84; 126, 194, 781 See Takahashi, K., 171; 194 See Thunell, R.C., 142, 170; 197 Honma, Y., 438; 467

Oceanography and marine biology

839

See Tamura, E., 426; 470 Hood, D.W. See Bader, R.G., 615; 746 See Naidu, A.S., 641; 771 See Slowey, J.F., 582; 780 Hood, J.M. See Ehrlich, K.F., 446; 465 Hoover, T.B., 600; 762 Hopkin, R., 542, 547; 762 Hopner, T., 593, 609; 762 Hope, H.-G., 86, 91, 92; 119, 121 See Iturriaga, R., 97; 122 Hori, K., 462; 467 See Haneda, Y., 497; 530 Horibe, Y. See Nozaki, Y., 184; 194 Horiguchi, Y. See Noda, H., 538; 773 Horn, B.M. See Horn, D.R., 624; 762 Horn, D.R., 624; 762 Horn, E.C., 538; 762 Horn, M.H., 420, 422, 427, 430, 433434, 443, 457; 467 See Buckley, J.L., 435; 463 Horne, R.A., 571, 640; 762 Horner, R., 94; 121 Horner, R.A., 1, 27, 28; 34 Horner, S.G., See Rowe, G.T., 114; 124 Hornung, H. See Roth, I., 636, 638, 643 643, 697, 701, 708, 710, 712; 778 Horridge, G.A. See Bullock, T.H., 402 403; 416 Horvath, G.J., 416, 601; 762 Hoshiai, T., 28; 35 Hoskins, K.D. See Beers, J.R., 558; 746 Hosokawa, I., 593; 762 Hoss, D.E., 553, 590, 718; 762 Hosteller, H.P., 554; 762 Houlihan, D.F., 394; 417 House, K.L. See Brown, D.H., 548; 749 Houston, M. See Soulsby, P.G., 587; 781 Houston, R.S., 375; 417 Howard, A.G., 550, 611, 612, 662, 677, 679; 762 Howarth, R.S., 552; 762 Howe, M.R., 38–73; 40, 44, 45, 47; 72 See Ambar, I., 42, 47, 50, 51, 53, 57, 58, 60, 63, 64, 67, 69, 70, 72; 72 Howell, M. See Williams, L.G., 715; 787 Howland, R.J.W. See Morris, A.W., 594; 771 Howmiller, R.P., 323; 343 Hrabĕ, S., 212, 291, 295; 343 Hrovat, D. See Orlob, G.T., 604; 774 Hsu, C.L. See Theis, T.L., 590; 784 Huang, C.P., 613; 762 See Elliott, H.A., 612, 601; 754 Hubbs, C.L., 487; 531

Interlinking of physical

1314

Hubschman, J.H., 565; 762 Huebner, J.D., 365, 394; 417 Hueck, H.J., 563, 630, 635, 655, 669, 690, 717, 730, 734; 762 Hueck-van der Plas, E.H., 538, 642; 762 Huggett, R.J., 578, 593; 762 See Bender, M.E., 569, 658; 747 See Bricker, O.P., 613; 748 See Helz, G.R., 617; 762 Huggins, A.K. See Boulton, A., 566; 748 Hughes, B.D., 210, 221; 343 See Learner, M.A., 209; 345 Hughes, D.J. See Mague, T.H., 78; 122 Hughes, G.M., 436; 466 Hughes, M.N., 537; 762 Huguenin, J.E., 563, 622; 762 Huizenga, D.L., 602; 762 Hulet, W.H., 489; 531 Hulings, N.C., 223, 260, 289, 336; 343 Huljev, D., 610; 762 Hume, D.N. See Westall, J.C., 608; 787 Hummel, B. See Delisle, C.E., 582; 753 Hummel, B.L., 491; 763 Hummerstone, L.G. See Bryan, G.W., 575, 588, 655, 658, 661, 665, 671, 677, 679, 725; 749 Humphris, S.E., 623; 763 Hung, T.C., 588, 616, 659; 763 Hunkins, K.E., 136; 193 Hunneman, D.H. See Eglinton, G., 89; 119 Hunt, D.T.E. See Foster, P., 614, 618; 757 Hunter, J., 208, 212, 214, 215, 221, 222, 228, 231, 276, 291, 295; 343 Hunter, J.R. See Postgate, J.R., 101; 124 Hunter, K.A., 594, 614; 763 See Marty, J.C., 122 Hunter, W.R., 566, 567, 718, 730; 763 Huntsman, S.A. See Barber, R.T., 560, 605; 745 See Smith, W.O., 97; 125 See Sunda. W.G.. 603, 620; 783 Huppert, J.E., 14, Fig. 16 (Tranter) facing p. 18; 35 Hura, M., 539; 763 Hurd, D.C. See Takahashi, K., 171; 195 Hurd, L.E. See Curtis, L.A., 358; 416 Huschenbeth, E. See Harms, U., 570; 760 Hussain, N.A., 422; 467 Hutchinson, T.C., 551, 606, 619, 742; 763 See Stokes, P.M., 548; 782 Hutner, S.H., 538; 763 Huynh-Ngoc, L. See Fukai, R., 636; 757 Hyman, L.H., 381, 402; 417 Hynes, H.B. N., 325; 343

Oceanography and marine biology Hyun, Y. See Fan, K.C., 162; 192 Ibragim, A.M., 542; 763 Ichikawa, T. See Nozaki, I.M., 469 Ichimura, S. See Seki, H., 91; 125 See Takahashi, M., 90, 92, 106; 126 Igarashi, H. See Kunimoto, M., 89; 122 Iizuka, S. See Hirayama, K., 543; 761 Ikeda, S. See Yamamoto, Y., 575; 788 Ikeda, T. See Reeve, M.R., 556; 777 Ikemura, K., 590, 623; 763 Ikonen, E. See Haka, P., 342 Ikuta, K., 584; 763 Ilus, E. See Bagge, P., 258, 270, 271, 287, 291, 293, 321, 323; 338 Imai, S. See Kusaka, Y., 766 Inamori, Y., 544, 725; 763 Innes, A.J. See Houlihan, D.F., 394; 417 Inoue, A., 636; 763 Inoue, S., 497; 531 See Shimomura, O., 497; 532 Inoue, Y. See Aoyama, I., 571; 745 Ireland, M.P., 578, 677, 680, 715; 763 Irvine, R. See Murray, J., 138; 194 Ishibashi, T., 455; 467 See Nishikawa, M., 454; 469 Ishii, T., 648, 650, 656, 661, 665, 698, 699, 704, 709; 763 See Yamamoto, Y., 575; 788 Ishiwatari, R., 88, 90; 121 Isibasi, M., 636; 763 Itoh, T. See Hayashi, M., 422, 423; 467 Ittekkot, V., 76; 121 Iturriaga, R., 82, 97; 121 See Morita, R.Y., 116; 123 Ivanova, A.A. See Konavalov, G.S., 629; 765 Ivanova, I.I. See Khovrychev, M.P., 541; 765 Ivanova, L.A. See Zhirova, V.V., 559; 789 Ivanovici, A.M., 408; 417 See Rainer, S.F., 407; 417 Ivlev, V.S., 217, 231; 343 Ivleva, J.V., 218, 233, 243, 276, 302, 317; 343 Iwai, T., 485, 492; 531 Iwamoto, T. See Hubbs, C.L., 487; 531 See Marshall, N.B., 487; 531 Izzo, G. See Melluso, G., 615; 770 Jackim, E., 568; 763 Jackson, G.A., 550, 604; 763 Jackson, R.W. See Rainbow, P.S., 574; 776

841

Interlinking of physical

1316

Jackson, W.B., 588; 763 Jacob, P.G. See Helmy, M.M., 544; 761 Jacobs, M.B., 133, 136, 139; 193 Jaffe, D., 641, 642, 643; 763 Jakobi, N. See Wiese, J., 325; 349 James, A.T. See Yaro, I., 89; 128 James, R.O. See Leckie, J.O., 593; 766 Jamieson, B.G.M., 206, 240, 274, 291; 343 See Brinkhurst, R.O., 198, 201, 208, 215, 319, 320, 322; 339 See Richards, K.S., 291; 347 Jan, T.-K., 587, 639, 664, 676, 682, 684, 686, 688, 696, 699, 704, 709; 763 See Hershelman, G.P., 586; 762 See Young, D.R., 589, 663; 788 Jannasch, H.W., 82, 87, 110, 111, 112, 115, 154; 121, 193 See Karl, D.M., 114; 121 See Tuttle, J.H., 114; 126 See Vaccaro, R.F., 90, 91; 126 See Wirsen, C.O., 110, 111; 127 Jansson, B.-O., 236, 237, 257, 260, 263, 267, 268, 276, 284, 288, 291, 299, 300, 301, 305, 306, 310, 334; 343 See Fenchel, T., 263, 276, 288, 334; 341 Jarman, N. See Brown, A.C., 389; 416 Jasper, S. See Hobbie, J.E., 86; 120 Jassby, A.D., 88; 121 Jedwab, J., 142, 169, 585; 192, 763 See Dehairs, F., 142; 191 Jefferies, D.F. See Cambray, R.S., 586; 750 See Dutton, J.W.R., 753 Jeffrey, L.M. See Eadie, B.J., 84; 119 See Slowey, J.F., 582; 780 Jehanno, C. See Lambert, C.E., 194 Jelinek, H. See Hartwig, E., 247, 248, 254; 342 See Pfannkuche, O., 245, 271, 296; 345 Jelnes, J. See Christensen, B., 217, 273, 312; 339 Jenkins, R.L., 625; 763 Jenne, E.A., 581, 601, 618, 629, 743; 763 See Ball, J.W., 603; 747 See Luoma, S.N., 617; 767 See Malcolm, R.L., 602; 768 Jennett, J.C., 560; 764 Jennings, J.B., 244; 343 Jensen, A., 548, 720, 722, 723; 764 See Braek, C.S., 551, 738; 748 See Eide, I., 554; 754 Jensen, J. See Christensen, B., 312; 339 Jensen, K. See Jørgensen, K.F., 544, 728, 739; 764 Jensen, M.J., 476; 531 Jerlov, N.G., 136, 518; 193, 531

Oceanography and marine biology Jerzmanska, A., 473, 515; 531 Jezequel, M. See Alayse-Danet, A.M., 539; 744 Jimenez, W. See Frick, L.P., 563; 757 Jitts, H.R. See El-Sayed, S.Z., 1; 34 Joensuu, O. See Bonatti, E., 614; 748 Johannes, R.E. See Webb, K.L., 78; 127 Johns, R.B., 89; 121 See Perry, G.J., 88; 124 See Volkman, J.K., 88; 126 Johnson, C.C. See Sick, L.V., 597; 780 Johnson, C.R., 457; 467 Johnson, D.L., 552, 613; 764 Johnson, F.H., 492, 493; 531 See Brown, F.A., 487; 529 See Haneda, Y., 492, 493; 529 See Shimomura, O., 494, 497; 532 See Sie, E.H.-C, 493; 532 Johnson, G.L. See Heezen, B.C., 63; 73 Johnson, G.W. See Boyd, C.M., 141; 191 Johnson, K.M. See Sieburth, J.McN., 98; 125 Johnson, M.G., 204, 231, 233; 343 Johnson, M.S., 430; 467 Johnson, P.G. See Villa, Jr, O., 643; 786 Johnson, P.W. See Sieburth, J.McN., 125 Johnson, R., 619; 764 Johnson, R.K. See Stephens, Jr, J.S., 446; 470 Johnson, R.L. See Ford, R.F., 756 Johnson, R.W., 88; 121 Johnson, S. See Griggs, G.B., 587, 643; 759 Johnson, S.R. See Hodge, V., 761 Johnson, V., 595, 640; 764 Johnston, D.B. See Waksman, S.A., 558; 786 Johnston, M.C. See Black, G.A.P., 558; 747 Johnston, N.T. See Northcote, T.G., 773 Johnston, W., 549; 764 Joint, I.R., 74–127; 91; 121 See Warwick, R.M., 233, 240; 349 Joiris, C., 99; 121 See Billen G., 90; 116 Jonasson, P. See Berg, K., 233; 338 Jonasson, P.M., 233; 343 Jones, A.H. See Burton, D.T., 569; 750 Jones, D., 422, 431; 467 Jones, D.P. See Laird, J.C., 193 Jones, E.J.W., 136; 193 Jones, G.B., 193, 618, 643; 764 Jones, G.E., 550, 551, 558, 583, 606, 645; 764 Jones, H.D., 385; 417

843

Interlinking of physical Jones, J.R.E., 567; 764 Jones, K.W. See Duedall, I.W., 754 Jones, L.H., 643; 764 Jones, L.P. See Gillespie, P.A., 91; 120 Jones, M.E. See Start, T.J., 722; 782 Jones, N.Y. See Cross, F.A., 575; 752 Jones, P.G.W. See Dutton, J.W.R., 754 Jones, R.L. See Hanson, H.C., 576; 760 Jordan, M.B. See Jones, G.B., 760, 618, 643; 764 Jordan, R.A. See Bender, M.E., 601; 747 Jørgensen, B.B., 105; 121 See Sørensen, J., 106; 125 Jørgensen, C.B., 75, 101, 313; 121, 343 Jørgensen, K.F., 544, 629, 635, 728, 740; 764 Jørgensen, N.O.G., 764; 764 Jørgensen, S.E., 637; 764 Joseph, E.B. See Saksena, V.P., 437; 470 Joshi, J.M. See Kappana, A.N., 764 Joyce, S.P. See Young, J.S., 789 Joyce, T.M., 29; 35 Joyner, T., 32, 585; 35, 764 Juget, J., 238; 343 Jumars, P.A. See Feller, R.J., 341 Jurinak, J.J. See Van Luik, A.E., 604; 786 Juste, C., 610; 764 Juteau, T. See Francheteau, J., 119, 757 See Speiss, F.N., 126 Kacprzac, J.L. See MacKay, N.J., 768 Kadar, S. See Brewer, P.G., 191 Kadko, D., 179; 193 Kadota, H. See Tanaka, N., 97; 126 Kähler, H.H., 297, 309, 313; 343 Kaibara, R. See Inoue, A., 763 Kain, J.M. See Hopkin, R., 542, 547; 762 Kajak, Z., 201; 343 Kakoi, H. See Inoue, S., 497; 531 Kalavati, C. See Subba Rao, B.V., 348 Kalf, J. See Eisma, D., 141, 151; 192 Kalin, J.R. See Gross, M.G., 759 Kalle, K., 136; 193 Kampa, E.M., 518; 531 Kampa, E.M. See Boden, B.P., 473, 518, See Young, R.E., 533 523; 529 Kamimura, S., 576, 715; 764 Kamp-Nielsen, L., 539; 764 See Steemann Nielsen, E., 474; 781 Kamykowna, B., 473; 531

1318

Oceanography and marine biology Kanazawa, K. See Kanazawa, T., 539, 541; 764 Kanazawa, T., 539, 541; 764 Kaneda, T., 88; 121 Kanias, G.D. See Papadopoulou, C., 663, 664, 668, 672; 774 Kanner, E. See Righi, G., 206; 347 Kantin, R. See Benon, P., 747 Kanwisher, J.W. See Maddux, W.S., 141; 194 Kaplan, I.R. See Brooks, R.R., 194; 749 See Claypool, G., 106; 117 See Goldhaber, M.B., 106; 119 See Nissenbaum, A., 106; 123 See Presley, B.J., 640; 776 See Simoneit, B.R., 780 See Sweeney, R.E., 84; 126 Kaplov, V.I. See Khobot’yev, V.G., 559; 764 Kappana, A.N., 636; 764 Karbe, L., 555, 726; 764 Karl, D.M., 88, 106, 115; 121 Karlson, P., 397; 417 Karrick, N.L., 539, 719, 739; 764 Kartar, S. See Hardisty, M.W., 713; 760 Kasprzak, K. See Dumnicka, E., 198; 340 Kastner, M. See Speiss, F.N., 126 Kasymov, A.G., 326; 343 Kates, M., 88; 121 Katsuragi, Y. See Miyake, Y., 194 Kattner, G.G., 107; 121 Katz, S., 611; 764 Katz, E.L. See Pringle, B.H., 689; 776 Kaufman, A. See Broecker, W.S., 191 Kaufman, Z.S., 327; 343 Kaushik, N.K. See Brinkhurst, R.O., 233, 237, 280; 339 Kauwling, T.J. See Reish, D.J., 777 Kawakami, Y., 643; 765 Kawano, S. See Werringloer, J., 539; 787 Kawashima, T. See Saijo, Y., 20; 35 Kazacos, M.N. See MacKay, N.J., 566; 768 Kazaikin, N.I., 586; 765 Kazarian, R., 765; 765 Kearney, J. See Stump, I.G., 783 Kearns, A. See Olafson, R.W., 574; 774 Keays, R.R. See Glasby, G.P., 624; 758 Keeney, D.R. See Dolar, S.G., 550; 753 Keeney, M., 544; 765 Keenleyside, M.H.A., 441; 467 Keillor, J.P., 591, 628; 765 Keller, D.M. See Matsumura, P., 109; 123 Kelly, D. See Kunz, Y., 426; 467

845

Interlinking of physical

1320

Kelly, M.G. See Tett, P.B., 473; 533 Kendall, Jr, A.W., 422, 448; 466 Kendall, D.R., 578; 765 Kendall, M.A., 210, 220, 226, 235; 343 Kendall, P.C. See Haven, D.S., 342 Kennedy, C.R., 215; 343 See Brinkhurst, R.O., 224, 230, 243, 258, 268, 280, 283, 291, 293, 320, 336; 339 Kenneth, J.P., 14; 35 Kenny, G.E. See Feller, R.J., 341 Kensley, B. See Penrith, M.-L., 422; 469 Kenyon, G.L. See Eberhard, A., 530 Kenyon, N.H., 63; 73 Keong, W.M. See Bok, C.S., 650, 656, 675, 678, 684, 693, 698, 707, 708; 748 Kepkay, P.E., 102, 122 See Novitsky, J.A., 102; 123 Kerkut, G.A., 545, 730; 765 Kerr, M.S. See Horn, E.C., 538; 762 Kerr, R.A., 601, 743; 765 Kerstitch, A.N. See Thomson, D.A., 422; 470 Kessel, M., 478, 483; 531 Kester, D.R. See Huizenga, D.L., 602; 762 Ketten, D.R. See Bishop, J.K.B., 83, 142; 117, 191 Kettling, R.C. See Colwell, R.R., 112, 114; 118 Kevern, N.R. See McIntosh, A.W., 540; 770 Key, G.S. See Stephens, Jr, J.S., 446; 470 Keyhani, E., 559; 765 Khailov, K.M., 612; 765 Khalid, R.A., 616, 643; 765 See Patrick, Jr, W.H., 601; 775 Khalily, H. See Afran, A., 642; 744 Kharkar, D.P., 617; 765 See Forster, W.O., 637; 756 See Turekian, K.K., 185; 197 Khobot’yev, V.G., 560, 603; 765 Khoo, H.W., 451; 467 Khovrychev, M.P., 541, 581, 619; 765 Khristoforoya, N.K., 658; 765 Kiauta, B. See Butot, L.J.M., 411; 416 Kiceniuk, J.W. See Fletcher, G.L., 538; 756 Kiener, A. See Casabianca, M.L.de, 445, 457; 464 Kilburn, R.N., 351, 358; 417 Kilham, P. See Kilham, S.S., 552; 765 Kilham, S.S., 552; 765 Kim, Y.H. See Elam, J.S., 462; 465 Kimmel, B.L., 154; 193 King, J.D., 88; 122 See White, D.C., 126 King, K.R., 100; 122

Oceanography and marine biology

847

King, L.H. See Rashid, M.A., 609; 776 King, M.J. See Fletcher, G.L., 538, 574, 575, 703, 713; 756 King, P.E. See Shackley, S.E., 442; 470 King, R. See Baier, R.E., 108; 116 Kinne, O., 304, 305, 311; 343 Kinney, P. See Williams, P.M., 111; 127 See Wilson, D.E., 609; 787 Kipphut, G. See Peng, T.H., 85; 124 Kirchgessner, M., 619; 765 Kirchner, W.P., 245; 344 Kiristayeva, N.M. See Barashkov, G.K., 746 Kirk, R., 217, 218; 344 Kirsch, R. See Nonotte, G., 436; 469 Kishpaugh, J.R.L. See Ali, S.A., 626; 744 Kislaliogu, M., 422, 456, 457, 459, 460; 466 Kitchen, J.C., 141; 193 Kitchens, W. See Gardner, L.R., 193; 757 Kjelleberg, S., 108, 109; 122 See Dahlbäck, B.O., 108; 118 Klapow, L.A., 744; 765 Klausewitz, W., 421; 467 Kleerekoper, H., 567; 765 Klein, D.H., 590; 765 Klein, E. See Spears, L.G., 622; 781 Klenova, M.V., 138; 193 Klima, J.R. See Merz, R.C., 612; 770 Kling, R., 560; 765 Kling, S.A. See Soutar, A., 195 Klinkhammer, G.P. See Bender, M.L., 614; 747 Kluczycka, K. See Zdybiewska, M., 622; 789 Klumpp, D.W., 650, 653, 654, 656, 668, 670, 680, 683; 765 Knauer, G.A., 81, 82, 83, 159, 552, 560, 578, 588, 637, 643; 122, 193, 765 See Boothe, P.N., 715; 750 Knauss, K.G. See Ku, T.L., 179; 193 Knezevic, M. See Chen, K.Y., 627; 750 Knezovich, J.P., 550, 732; 765 Knights, P.J. See Evans, P.R., 244; 341 Knittel, M.D. See Hetrick, F.M., 546; 761 Knöller, F.H., 197, 257, 268, 269, 271, 272, 276, 280, 287, 291, 320, 321; 343 Knook, D.L. See Spronk, N., 563; 781 Knox, G.A., 1, 5; 35 Kobayashi, H. See Nozaki, I.M., 469 Kobayashi, N., 557; 765 Kobayashi, T., 438, 440, 448, 459; 467 Koch, F.A., 588; 765 See Benedict, A.H., 591; 747 Koefoed, J.H. See Hesthagen, I.H., 425; 467 Koeller, P., 568, 713, 715, 743; 765

Interlinking of physical

1322

Koene, H., 209, 210, 226, 230, 233, 235; 344 Koide, M. See Bruland, K.W., 191 See Goldberg, E.D., 587; 758 See Williams, P.M., 84; 127 Koike, I., 103, 105; 122 See Seki, H., 110; 125 Kojanayi, T. See Ishii, T., 648; 763 Kölliker, A., 249, 253, 254; 344 Kolodny, Y. See Presley, B.J., 640; 776 Komar, P.D., 163, 166; 193 Konavolov, G.S., 629; 765 Kondo, N. See Hosokawa, I., 593; 762 Kononova, M.M., 608; 766 Konovalova, I.W. See Sorokin, Y.I., 98; 126 Kopfler, F.C., 575; 766 Kornberg, H.L., 99; 122 Kornicker, L.S., 355, 496; 417, 532 Korringa, P., 538, 581, 742; 766 Kosiorek, D., 215; 344 Kosman, D.J. See Shatzman, A.R., 537; 780 Kotrschal, J. See Goldschmid, A., 457; 466 Kotsaki-Kovatsi, V. See Kovatisis, A., 766 Kounaves, S.P. See Zirino, A., 608; 789 Kovacs, M., 576; 766 Kovatisis, A., 668; 766 Koyama, T. See Matsuda, H., 89; 123 Koyanagi, T. See Nakahara, M., 713; 772 Kramer, C.J.M. See Duinker, J.C., 593, 635; 754 Kramer, J.R. See Van den Berg, C.M.G., 605; 785 Kraner, H.W. See Duedall, I.W., 754 Krauter, K. See Stokes, P.M., 548; 782 Krautgartner, W.-D. See Goldschmid, A., 457; 237 Krauskopf, K.B., 602, 617, 618, 628; 766 Kremling, K., 617, 635, 638; 766 Kretzschmar, J.G., 586; 766 Krey, J., 138; 193 Krieger, N. See Morin, J.G., 532 Krijgsman, B.J., 371; 417 Krishnamurthy, K. See Sundararaj, V., 594; 783 Krishnan Nambisan, P.N.N. See Lakshmanan, P.T., 563, 606, 735; 766 Krishnaswami, S., 82, 137, 142, 160, 177, 181, 184, 185; 122, 193 See Bhat, S.G., 190 See Craig, H., 180; 191 See Spencer, D.W., 84; 126, 194, 781 Kriss.A. E., 109; 122 Krizenecky, J., 297; 344 Kronfeld, J. See Navrot, J., 773 Kroopnick, P. See Craig, H., 84; 118

Oceanography and marine biology Kroopnick, P.M., 84; 122 Krüger, F., 230, 305, 317, 319; 344 Krzyz, R.M. See Saliba, L.J., 549, 564, 739; 778 Ku, T.L., 179; 193 Kühn, M. See McLachlan, A., 352; 417 Kuipers, B., 424, 448, 459, 461; 467 Kumaraguru, A.K., 562, 732; 766 Kunimoto, M., 89; 122 Kunz, Y., 426, 428; 467 Kupryszewski, G. See Bojanowski, R., 605; 748 Kurata, A., 559, 643, 720; 766 Kurata, K. See Isibasi, M., 763 Kurihara, Y. See Inamori, Y., 544, 725; 763 Kurpyakova, Z.N. See Petrov, Y.M., 775 Kurt, L.A., 237; 344 Kusaka, Y., 636; 766 Kushner, D.J. See Laube, V., 604; 766 Kusuma, M.S. See Neelakantan, B., 643; 773 Kusumgar, S. See Somayajulu, B.L.K., 176; 195 Kuwabara, S., 483; 532 Laakso, M., 227, 269, 271, 287, 291, 293; 344 Labeyrie, L.D., 176, 179; 193 Lackie, N. See Erickson, S.J., 554; 755 Lacombe, H., 38, 57, 61; 74, 73 Ladle, M., 208, 221; 344 Lagerwerff, J.V., 743; 766 Lagios, M.D. See McCosker, J.E., 480; 532 Lagrange, R. See Bardet, J., 746 Laing, W.A. See Christeller, J.T., 84; 118 Laird, J.C. 160; 194 Lakshmanan, P.T., 563, 606, 735; 766 Lal, D., 130, 136, 141, 150, 168, 171, 174, 176; 193 See Bhat, S.G., 190 See Krishnaswami, S., 137, 142; 191 See Slinn, W.G. N., 780 See Somayajulu, B.L.K., 175; 194 Lamb, C.A., 536; 766 Lambert, C.E., 142, 147, 163, 169, 172; 194 Lampert, W., 78; 122 Landau, J.V., 109; 122 See Pope, D.H., 110; 124 See Smith, W., 110; 125 See Swartz, R.W., 110; 126 Lande, E., 643, 666, 679, 683, 686, 693, 694, 696, 698, 699; 766 Landing, W.M., 649; 766 See Bruland, K.W., 158; 191 Lane, C.E. See Bernard, F.J., 538, 545, 565; 747

849

Interlinking of physical

1324

Lang, J.M. See Bell, W.H., 101; 117 Länge, R., 235; 343 Langille, W.M. See Young, E.G., 577; 789 Langon, M. See Juste, C., 610; 764 Lankaster, E.R., 246; 344 Lännergren, C., 81; 122 Lansford, L.M. See Marvin, K.T., 626; 769 Lantz, J.B., 587; 766 Lantzy, R.J., 586; 766 Lapan, Jr, R.L. See Eisler, R., 755 Larcombe, R.A. See Badcock, J., 519; 528 LaRoche, G., 554; 766 See Gardner, G.R., 546, 568; 756 Larrance, J.D., 156, 159; 194 Larrick, S.R. See Cherry, D.S., 577; 751 Larsen, P.F., 563, 665, 689; 766 Larson, R. See Speiss, F.N., 126 Larsson, U., 94, 97; 122 La Rue, G.R., 203; 343 Lasiak, T.A. See McLachlan, A., 417 Lasik, Y.A., 611; 766 Lassèrre, P., 197, 198, 199, 202, 203, 208, 219, 223, 224, 225, 226, 228, 230, 237, 247, 257, 258, 263, 264, 266, 269, 275, 280, 288, 289, 290, 299, 300, 312, 313, 314, 316, 318, 319; 344 See Erséus, C., 197, 198, 287, 290; 340 Latimer, I., 135; 194 Latogurskii, V.I., 30; 35 Latter, P.H., 204; 344 Latz, M.I. See Warner, J.A., 523; 534 Laube, V., 604, 641; 766 Lauder, G.V. See Spender, D.L., 459; 470 Laumen, J., 439; 468 Laumond, F. See Aubert, M., 745 See Gnassia-Barelli, M., 559; 757 Lavenburg, R.J. See Fitch, J.E., 420; 465 Laver, M.B. See Smith, K.L., 113; 126 Laverack, M.S. See Bailey, D.F., 399; 415 Lavoie, D.M. See Sieburth, J. McN., 98; 125 Lavrov, V.M. See Klenova, M.V., 193 Lawry, J.V., 509, 523, 525; 532 Laws, R.M., 1, 6, 8; 35 Lazareva, E.A. See Romanov, A.S., 637; 778 Leadbetter, E.R., 116; 122 Lear, D.W. See Palmer, H.D., 626, 646, 687; 774 Learner, M.A., 210, 217, 218, 325; 344 Leaseburge, C. See Fan, K.C., 162; 192 Le Blanc, M.J. See Andersen, R.J., 611; 744 Lebour, M.V., 375; 417 Lechevalier, M.P., 88; 122

Oceanography and marine biology

851

Leckie, J.O., 593, 604; 766 See Davis, J.A., 766, 612, 614, 615, 618; 752 See Lion, L.W., 603; 767 LeCorroller, Y. See Floch, H., 563, 739; 756 Lee, C. See Hobbie, J.E., 240; 343 Lee, C.C. See Bunt, J.S., 27; 33 Lee, G.F., 554, 591, 719; 766 See Gjessing, E.T., 608; 757 See Lopez, J.M., 618; 767 See Sanchez, I., 564; 778 Lee, H.T. See Murphy, L.S., 548; 772 Lee, J.C. See Chang, K.-H., 422, 457; 464 See Ryu, B.S., 424; 470 Lee, J.J., 112; 122 See Tietjen, J.H., 204; 348, 349 Lee, R.F. See Gardener, W.S., 333; 341 Lee, S.-C, 422, 423, 442, 457; 468 See Chang, K.-H., 422, 448, 456; 464 Lee, T. See Craig, Jr, H.L., 752 Leedow, M.I. See MacKay, N.J., 567; 768 Lehman, D.S., 538, 607; 766 Lehman, J., 540, 717; 766 Lehman, J.L., 540, 717, 721; 766 Lehmann, F.E., 203; 344 Lehnen, O. See Bergmann, M., 593; 747 Lehner, C.E. See Thomson, D.A., 421, 422, 423, 431, 449, 450, 457; 470 Lehninger, A.L., 581; 766 Leiberman, S.H., 606, 613; 766 Leisma, A. See Haka, P., 342 Leisman, G., 483, 489; 532 See Nealson, K.H., 532 Leland, H.V., 536, 571, 600, 629, 642, 644, 646, 680, 682, 683, 733, 738; 766 Leloup, E., 268; 344 Lemke, A.E. See Helmy, M.M., 544; 761 Lennarz, W.J., 89; 122 Lennox, Jr, T.J. See Groover, R.E., 590; 758 Lense, A.H. See Jenkins, R.L., 625; 763 Leo, R.F., 87, 88; 122 Leon, B. See Brown, A.C., 352, 404; 416 Leonard, J.D. See Rashid, M.A., 608; 776 Le Pichon, X. See Ewing, M., 135; 192 Leppäkoski, E., 271, 321, 322, 334; 344 Leppäkoski, E.J., 326, 332; 344 Le Ray, W., 203; 344 Lerch, K., 574; 767 Lerman, A., 130, 134, 141, 150, 153, 164, 165, 168, 172, 174, 180; 194 See Lal, D., 151, 171; 193 Lessel, E.F. See Buck, J.B., 475; 529

Interlinking of physical

1326

Leumer, G., 590; 767 Leveau, M. See Benon, P., 747 Lever, A.J. See Thijssen, R., 459; 470 Lev-er, J., 589, 642; 766 Lever, J. See Thijssen, R., 459; 470 Le vine, W.G., 606; 766 Levy, C. See Tamblyn, N., 622; 784 Lewis, A.G., 535–789; 538, 543, 550, 556, 560, 562, 568, 578, 583, 600, 616, 628, 714, 743; 767 See Whitfield, P.H., 550, 552, 604; 787 Lewis, C.W. See Baptist, J.P., 714; 746 Lewis, D.B., 421; 468 Lewis, J.A.M. See Sunda, W.G., 550, 551, 603; 783 Lewis, J.M. See Cross, F.A., 752 Lewis, J.P.B., 381; 417 Lewis, R.H. See Klapow, L.A., 743; 765 Lewis, S.D., 546; 767 Lewis, W.M. See Lewis, S.D., 546; 767 Leyden, D.E., 636; 767 Li, M.W. See Wong, M.H., 552, 635, 643, 677; 788 Li, Y.-H., 180, 185; 193 See Hung, T.C., 587, 658; 762 Lichtfuss, R., 588, 597; 767 Lieberman, S.H., 607, 613; 767 See Zirino, A.H., 601, 612; 789 Liebezeit, G., 76; 122 See Meyer-Reil, L.-A., 238; 345 Lieser, K.H. See Heuss, E., 608; 761 Likens, G.E. See Eaton, J.S., 162; 192 Lillie, F.R., 538, 566, 727, 740; 767 Lilly, D.M. See Hanna, B.A.,498; 760 Limberger, D. See Taborsky, M., 432, 456, 458; 470 Lindén, O., 328; 344 Lindley, J.A., 78; 122 See Williams, R., 78; 122 Lindquist, D.G., 453, 457; 468 Lindsay, W.L., 550; 767 Lindström, L.S. See Leppäkoski, E.J., 326, 332; 344 Linke, O., 258, 264, 267; 344 Linley, A.E.S. See Newell, R.C., 94; 123 Linthicum, D.S. See Tebo, B.M., 484; 533 Lion, L.W., 604; 767 Lisitzin, A.P., 133; 194 Lisk, D.J. See Young, R.G., 560; 789 Liss, P.S., 107; 122 See Emerson, S., 618; 754 See Slinn, W.G.N., 780 Litchfield, C.D., 76, 106; 122 Livingstone, H.D. See Labeyrie, L.D., 179; 193

Oceanography and marine biology Lochhead, G, See Learner, M.A., 210; 344 Locket, N.A. See Herring, P.J., 519, 522; 531 Lockwood, A.P.M., 539; 767 Lockwood, S.J., 422, 424, 448, 459; 468 Loden, M.S., 199, 244, 284; 344 Loew, E.R., 426; 468 Lohmann, H., 245; 345 Long, D.T., 604; 767 Lonsdale, P., 115; 122 Lontie, R., 612; 767 Loosanoff, V., 203; 345 Lopez, J.M., 590, 618; 767 See Lee, G.F., 591; 766 Lorenzen, C.J. See Copping, A.E., 79, 100; 118 Loretz, C.A., 429; 468 Loring, D.H., 539, 588, 594, 616, 642, 643; 767 Lorz, H.W., 545, 592, 736; 768 See Schreck, C.B., 544, 558; 780 Losey, Jr, G.S., 439, 440, 452; 468 Loutit, M. See Patrick, F.M., 325, 558, 715; 346, 775 Loutit, M.W. See Patrick, F.M., 558, 579; 775 Lovelock, R.L. See Cardeilhac, P.T., 750 Lowenstine, M. See Case, J.F., 523; 529 Lowman, F.G., 594, 595, 636, 639, 649; 768 Lowry, J.K. See Knox, G.A., 1; 35 Lowthio, D. See Soulsby, P.G., 587; 781 Lu, J.C.S., 613, 643; 767 Lubbock, R., 421, 422, 423; 468 Lucas, C.E., 578; 768 Lucas, M.I. See Newell, R.C., 94; 123 Luciano, D. See Phelps, D.K., 577; 775 Lum-Shue-Chan, K. See Chau, Y.K., 602; 750 Lunde, G., 577, 642, 651, 656; 768 Lundkvist, H., 203; 345 Lunt, O.R. See Alexander, G.V., 744 Lunz, J.D., 551, 576, 603; 768 Luoma, S.N., 551, 581, 603, 616, 621, 659; 767 See Jenne, E.A., 582, 601; 763 See Leland, H.V., 536, 570, 601; 766 Luther, A., 245; 345 Luther, G. See Ramachandran Nair, P.V., 422; 470 Luxon, P.L. See Wong, P.T.S., 742; 788 Luyendyk, B. See Speiss, F.N., 126 Lynch, R.V. See Tsuji, F.I., 494; 534 Lyons, W.B., 626, 642; 768 Lythgoe, J.N., 426; 468 See Loew, E.R., 426; 468

853

Interlinking of physical

1328

MacArthur, D.M., 622; 768 MacCarthy, P., 604; 768 Macaulay, I.D. See Bewers, J.M., 768; 747 MacDonald, C.K., 421, 422, 424, 439, 443, 457, 461; 468 MacDonald, D.D. See Syrett, B.C., 590; 784 MacDonald, K.C. See Speiss, F.N., 126 Macdonald, R.W, See De Mora, S, J., 623; 753 MacDonald, S.M. See Mason, C.F., 461; 468 MacDougall, J.D. See Speiss, F.N., 126 MacFarlane, R.B., 608; 768 Machemer, L. See Blüm, V., 437; 464 MacInnes, J.R., 407, 544, 547, 551, 562, 619, 620, 718; 417, 768 MacIntyre, F., 107; 122 MacIsaac, J.J. See Barber, R.T., 605; 746 MacKay, D.W. See Halcrow, W., 760 MacKay, N.J., 567, 659; 768 See Bebbington, G.N., 747 MacKenzie, F.T, See Lantzy, R.J., 586; 766 Mackie, G.O., 490; 532 Mackintosh, N.A., 3, 15, 29; 35 MacNae, W., 422, 425, 435, 440, 448, 462; 468 Madan, M. See Thind, K.S., 537; 784 Maddux,W. S., 141; 194 Madelain, F., 38, 41, 42, 44, 59, 63, 71; 73 See Lacombe, H., 38, 60; 73 Maeda, M., 602; 768 Maestrini, S.Y. See Berland, B.R., 554; 747 Magazzu, G. See Calapaj, R., 750 Magee, R.J. See Talbot, V., 573, 613; 784 Maghagen, C. See Edlund, A.-M., 459; 465 Mague, T.H., 77; 122 Magnus, D.B.E., 439, 452, 456, 459; 468 Magnuson, V.R., 601; 768 Magnusson, B., 634, 637; 768 Mainwaring, D.J. See Ehrlich, H.L., 623; 754 Major, C.W. See Scott, D.M., 563, 734; 780 Majori, L., 564, 566, 573; 768 Majtan, V., 101; 122 Makarov, R.R., 29; 35 Makemson, J.C., 477; 532 Malcolm, R.L., 602; 768 Malloy, S.C. See Barvenik, F.W., 91; 117 Maloney, T.E. See Erickson, S.J., 554; 755 Mamayev, O.I., 59; 73 Mamen, C., 626; 769 Manabe, T. See Hiraizumi, Y., 642; 761 Manahan, S.E., 538, 603; 769 Mancinelli, G. See Viarengo, A., 786

Oceanography and marine biology

855

Mancy, K.H., 554; 769 Mandel, M. See Jensen, M.J., 531 Mandelli, E.F., 540, 552, 560, 562, 589, 718, 721, 723, 731, 732; 769 See Burkholder, P.R., 1, 27, 28; 33 See El-Sayed, S.Z., 1; 33 See Zeitoun, M.A., 589; 789 Mangini, A. See Dominik, J., 588; 753 Mangum, C.P., 317, 320; 345 See Weiland, A.L., 536; 785 Manheim, F. See Meade, R.H., 123 Manheim, F.T., 134, 137, 138; 194 Manley, A. See Davenport, J., 556, 717; 752 Manley, A.R., 543, 563, 717, 732, 733, 739; 769 Manning, M.J. See McCapra, F., 498; 532 Manning, P.G., 608; 769 Mansuri, A.P. See Bhan, S., 428; 464 Mantoura, A.J. See Morris, A.W., 594; 771 Mantoura, R.F.C., 583, 602; 769 Manuels, M.W., 88; 118 Manwell, C., 319; 345 Marchig, V. See Halbach, P., 624; 760 See Schnier, C., 624; 780 Marcus, E., 198, 205, 206; 345 Marechal, G. See Baguet, F., 510, 511, 512; 528 See Christophe, B., 511; 529 Mares, A. See Gächter, R., 550; 757 Margolis, S.V. See Dugolinsky, B.K., 624; 754 Mariani, G.M. See Lee, G.F., 591; 766 Markov, G.S. See Andreev, V.V., 571, 713; 745 Marks, G.W., 562, 734; 769 Marley, J.J. See Theis, T.L., 590; 784 Marliave, J.B., 427, 434, 437, 441, 446, 448, 462; 468 Marquenie-Van der Werff, M. See Ernst, W.H.O., 555, 576; 755 Marquis, R.E., 109; 122 See Matsumura, P., 110; 123 Marr, J., 15, 30; 35 Marsh, B., 422, 423, 445, 450, 454, 461; 468 Marshall, D.W., 622; 769 Marshall, N.B., 425, 487, 515, 518, 526; 468, 532 Marshall, W.S., 429; 468 Marsland, D.A. See Brown, F.A., 487; 529 Martell, A.E., 606; 769 Martens, C.S., 106; 123 Martin, C.L. See Martin, R.A., 440; 468 Martin, D.F., 543, 610; 769 Martin, J.H., 555, 580, 649, 662, 663, 668, 715; 769 See Anderlini, V.C., 569; 744 See Flegal, A.R., 550; 755

Interlinking of physical

1330

See Goldberg, E.D., 758 See Knauer, G.A., 81, 158, 159, 561, 578, 637; 122, 193, 765 See Martin, M., 552; 768 Martin, J.L.M., 673; 769 Martin, J.M., 592; 769 See Reish, D.J., 548, 586; 777 Martin, M., 553, 589, 717, 734; 769 Martin, M.J., 543, 563, 734; 769 Martin, R.A., 440; 468 Martin, S.G., 563; 769 Martin-Neumann, U. See Muller, R., 429; 469 Martoja, M., 585; 769 Marty, J.C., 107; 123 See Barbier, M., 107; 116 Marusic, E.T., 433, 434, 435; 468 Marvin, K.T., 626; 769 Mason, C.F., 221, 461; 345, 468 Massé, H. See Bodoy, A., 309; 338 Massie, K.S. See Vivian, C.M.G., 588; 786 Massie, L.C. See Litchfield, G.D., 76; 122 Massoth, G.J. See Landing, W.M., 586; 766 Mast, S.O. See Bertholf, L.M., 538; 747 Mathemeier, P.F. See Morita, R.Y., 109; 123 Mathews, C.W. See O’Neill, T.B., 623; 773 Mathews, T.D., 659; 769 Mathieu, G. See Hunkins, K.E., 136; 193 Mathieu, G.G. See Ku, T.L., 179; 193 Mathis, J.H., 579; 769 Matis, J.A. See Kleerekoper, H., 567; 765 Matson, W.R. See Bender, M.E., 601; 747 Matsuda, O. See Taga, N., 86, 87; 126 Matsuda, H., 89; 123 Matsui, K. See Enomoto, N., 640; 755 Matsui, Y., 642; 769 Matsuike, K., 24; 35 Matsumoto, E., 177; 194 Matsumoto, T., 660, 673, 688; 769 Matsumura, P., 109; 123 See Marquis, R.E., 110; 122 Matte, A. See Waldhauer, R., 742; 786 Matthiae, A., 294; 345 Mattice, J.S., 590; 769 Maurer, D. See Haines, J.L., 267; 342 Maurer, L. See Parker, P.L., 88; 124 Maxwell, J.R. See Brooks, P.W., 117 Mayberry, W.R. See Fazio, S.D., 88; 119 Mayer, F. See Romesser, J.A., 125 Mayer, J. See Kopfler, F.C., 575; 766

Oceanography and marine biology

857

Mayer, R.F., 432, 439, 457; 468 Maynard, S.D. See Young, R.E., 534 Maynard, V. See Bender, M.L., 117 Mazurek, M.A. See Simoneit, B.R., 780 McAllister, D.E., 525; 532 McArthur, J.M., 623; 769 McBrien, D.C.H., 539; 769 McCall, P.L., 281, 322; 345 McCance, R.A., 675, 677, 679; 769 McCapra, F., 497, 498; 532 McCarthy, R. See Gordon, M.S., 466 McCartney, M.J. See Calvert, S.E., 131; 191 McCartney, M.S., 21, 26; 35 McCauley, R.N., 326; 345 McCave, I.N., 82, 141, 147, 150; 123, 194 McClin, R. See Lowman, F.G., 768 McComas, F.T. See Thompson, J.A.J., 591, 626, 642, 643; 784 McConnaughey, F., 480, 481, 491; 532 McCormick, S.J. See Anderlini, V.C., 745 McCosker, J.E., 480, 481, 483; 532 See Stephens, Jr, J.S., 445; 470 McCrady, J.K., 608; 769 McDermott, D.J., 567, 587, 691; 769 See Alexander, G.V., 744 See Young, D.R., 591 McDermott-Ehrlich, D., 682, 709; 769 See Jan, T.-K., 763 See Young, D.R., 788 McDuffie, B., 592, 642; 770 McElhone, M.J., 209; 345 McElroy, W.D. See Sie, E.H.-C, 493; 533 McEntyre, C.L. See Minear, R.A., 771 McGowan, J.A. See Enright, J.T., 115; 119 McGowen, R.E. See Ehrlich, K.F., 446; 465 McGrath, M.S., 588, 635; 770 McGrath, R. See Greig, R.A., 643; 759 McGreer, E.R., 544, 562, 587, 644, 659, 665, 718; 770 See Chapman, P.M., 588; 750 McGwynne, L.E., 352, 355, 357, 358, 362, 365, 366, 369, 370, 371, 372, 373, 374, 377, 389, 394, 411; 416 See Dye, A.H., 354, 388, 390, 394; 416 See McLachlan, A., 417 McHugh, D.J. See Brooks, P.W., 117 McIlhenny, W.F. See Mandelli, E.F., 552, 589, 731, 732; 769 See Zeitoun, M.A., 589; 789 McInerney, M.J., 105; 123 McInnes, J.R. See Calabrese, A., 750 McIntosh, A. See Peyton, T., 585; 775

Interlinking of physical

1332

McIntosh, A.W., 540; 770 McIntyre, A.D., 201, 225, 236, 336, 539, 648; 345, 770 See Munro, A.L.S., 334; 345 McKenzie, R.M., 625; 770 McKee, J.E., 540, 742; 770 McKim, J.M., 569; 770 McKinley, P.W. See Malcolm, R.L., 602; 768 McKnight, D.M., 607, 612; 770 See Swallow, K.C., 783 McLachlan, A., 201, 206, 230, 260, 261, 271, 276, 278, 288, 336, 352, 355, 357, 367, 368, 369, 374, 395; 345, 417 See Ansell, A.D., 354, 370, 371; 414 McLeese. D.W., 720; 770 McLerran, C.J., 624; 770 McLusky, D.S., 226, 233, 258, 264, 300, 321, 322, 326, 334, 562, 717, 719; 345, 770 See Ansell, A.D., 369; 619 See Bagheri, E.A., 219, 221, 223, 225, 226, 228, 231, 233, 235, 264, 325; 338 McManus, D.A. See Sternberg, R.W., 195 McNally, P.M. See Carlucci, A.F., 103; 118 McNeilly, T. See Gartside, D.W., 548; 758 McNulty, J.A., 525; 532 McPherson, B.P. See Lorz, H.W., 546, 592, 736; 768 McWhinnie, M.A. See Rakusa-Suszczewski, S., 9; 35 McWilliams, D.C. See Oliff, W.B., 346 Meaburn, G.M. See Hall, R, A., 644; 760 Mead, G.W. See Backus, R.H., 528 Meade, R.H., 82; 123 See Manheim, F.T., 133, 138; 193 Meadows, P.S., 236, 283; 345 Means, J.L., 605; 770 Mears, H.C., 700, 706, 711; 770 See Eisler, R., 754 Mearns, A.J. See Alexander, G.V., 744 See Bascom, W., 747 See De Goeij, J.J.M., 752 See McDermott, D.J., 567; 769 See Reish, D.J., 777 See Young, D.R., 664, 681, 686, 694, 696, 701, 703, 704, 708, 711, 712, 714; 788 Meeter, D.A. See White, D.C., 127 Meguro, H., 1, 27, 28; 35 Meier, M., 247, 253; 345 Meijer, C.L., 739; 770 Mein, B., 227; 345 Meincke, J. See Gieskes, J.M., 40; 72 Meineke, T., 223, 259, 260; 345 Meith, S.J. See Hazel, C.R., 569, 736; 761 Melhuus, A., 551, 554, 635, 645, 652, 653, 654; 770 Melluso, G., 615, 689; 770

Oceanography and marine biology Mel’Nik, P.I. See Kazaikin, N.I., 586; 765 Melsom, S. See Eide, I., 554; 754 See Myklestad, S., 561, 575; 771 Mencher, F.M. See Young, R.E., 522; 534 Menon, M.P., 595; 770 Menzel, D.W., 110, 592, 713, 743; 123, 770 See Goering, J.J., 107; 119 Menzies, D. See Pak, H., 136; 194 See Kitchen, J.C., 142; 192 Menzies, D.W. See Zaneveld, R.J.V., 136; 195 Meon, A.N., 393; 417 Mercer, B.S. See Serne, R.J., 636, 644; 780 Meredith, F.L., 352, 365; 417 See Brown, A.C., 354, 371, 392, 394, 406; 416 Mericam, P., 600; 770 Mero, J.L., 625; 770 Merrett, N.R., 491; 532 Merrill, A.S. See Wensloff, D.R., 571; 787 Merwin, E.A. See Miller, W.E., 555; 771 Merz, R.C., 612; 770 Mesnil, F. See Caullery, M, 254; 339 Metcalf, P.J. See Goulder, R., 559; 759 Meulen, H., 635; 770 Meurs, C.J. See Van Der Weijden, C.H., 592; 786 Meybeck, M. See Martin, J.M., 592; 769 Meyer, R.L., 542; 770 Meyer-Reil, L.-A., 236, 237, 240; 345 See Faubel, A., 239; 340 Meyer-Rochow, V.B., 481; 532 Meyers, P.A., 614; 770 Miall, L.M. See Armstrong, E.F., 621; 745 Michaelis, W. See Ittekkot, V., 76; 121 Michaelsen, W., 205, 238, 263, 268, 269, 271, 287, 291, 323; 345 Michel, H. See Craig, Jr, H.L., 752 Michel, J. See Moore, W.S., 185; 194 Michelson, A.M. See Henry, J.P., 477; 531 Michnowsky, E. See Chapman, P.M., 322; 339 Middlebrooks, E.J. See Filip, D.S., 580; 756 Mie.G., 134; 194 Mieschner, D. See Paul, J., 641; 775 Milanovich, F.P., 550, 579, 658; 770 Milbrink, G., 199, 321, 323; 345 Milburn, H.B. See Larrance, J.D., 159; 194 Miles, P.S., 452, 453; 468 Mill, P.J., 312; 345 Mills, A.L. See Austin, B., 558; 745 Mills, B. See Bowser, C.J., 624; 748 Miller, D., 76; 123

859

Interlinking of physical

1334

Miller, J.C. See Biggs, R.B., 633; 770 Miller, J.E. See Calabrese, A., 750 Miller, J.M. See Duce, R.A., 754 Miller, M.A., 558, 726; 770 Miller, O. See Leloup, E., 268; 344 Miller, P.J., 442, 457; 468 See Fouda, M.M., 442; 465 Miller, S. . See Speiss, F.N., 126 Miller, W.E., 555; 770 Millero, F.J., 552, 638; 771 Milton, P., 423, 427, 428; 354 See Dixon, P.N., 423; 464 Minagawa, M. See Tsunogai, S., 181; 195 Minear, R.A., 600; 771 Mitchell, G. See Hastings, J.W., 486; 531 Mitchell, R., 559, 562; 771 See Bell, W.H., 101; 116 See Chet, I., 561; 750 Mitchell, R.L. See Black, W.A.P., 576, 635; 747 Mitropolskii, A.Yu. See Babinets, A.E., 746 Mitskevich, I.N. See Kriss, A.E., 109; 122 Miyagawa, K. See Nozaki, I.M., 469 Miyake, Y., 177; 194 See Sugimura, Y., 601; 782 Mizushima, Y., 540; 771 Möbius, K., 248; 345 Mock, Jr, W.R. See Fox, L.S., 422; 465 Modonutti, G.B. See Majori, L., 566; 768 Moffatt, N.M., 438; 468 Mohankumar, K.C., 109; 123 Mohri, H. See Morisawa, M., 573; 771 Mohus, A. See Braek, C.S., 551, 739; 748 Mollick, R.S., 445, 447, 453; 468 Montgomery, J.R., 577, 587, 593, 605, 614, 636, 715; 771 Montgomery, W.L. See Rosenblatt, R.H., 480, 481, 483; 533 Montgomery, W.M., 457; 468 Mook, W.G. See Salomons, W., 641, 643; 778 Mooney, J.R., 619; 771 Moore, C., 644; 771 Moore, J.R. See Morgan, C.L., 624; 771 Moore, J.W., 235; 345 Moore, M.D. See Bascom, W., 746 See Jan, T.K., 586; 763 See Young, D.R., 589; 788 Moore, R.M., 788, 635; 771 Moore, W.S., 184, 185; 194 See Bhat, S.G., 190 See Sarmiento, J.L., 194

Oceanography and marine biology

861

Moosleitner, H., 437; 468 Mopper, K. See Degens, E.T., 76; 119 Moraes, C.F. See Zingde, M.D., 789 Moraitou-Apostolopoulou, M., 548, 565, 720, 727; 771 Morales, E., 685; 771 Morel, F.M.M. See Andersen, D.M., 538, 543, 551, 559, 603, 610, 717, 722, 738; 745 See Anderson, M.A., 602; 744 See McKnight, D.M., 608, 611; 769 See Morel, N.M.L., 540; 770 See Swallow, K.C., 617; 783 See Westall, J.C., 551, 608; 787 Morel, N.M.L., 538, 541; 770 See Swallow, K.C., 783 Morell, A.G. See Scheinberg, I.H., 612; 779 Morgan, C.L., 624; 771 Morgan, D. See Pesch, C., 550, 614, 724; 775 Morgan, J.J. See Jackson, G.A., 550, 604; 763 See Stumm, W., 617; 782 See Vuceta, J., 618; 785 Morgan, M.J. See Davey, E.W., 560; 752 Morgan, W.S.G., 557; 771 Moriarty, D.J.W., 88; 123 Morii, H., 434; 468 See Tamura, S.O., 435; 470 Morin, J.G., 480, 481; 531 See Herring, P.J., 473, 480, 483, 485, 486 487, 503, 511, 515, 523; 531 See Ruby, E.G., 476, 487, 488, 490; 532 Moring, J.R., 420, 421, 422, 423, 424, 425, 443, 450, 462; 469 Morisawa, M., 573; 771 Morita, R.Y., 109, 110, 114, 116; 123 See Baross, J.A., 110; 116 See Gillespie, P.A., 90; 119 See Novitsky, J.A., 101, 110; 123 See ZoBell, C.E., 114; 127 Morita, Y., 636; 771 Moroz, T.G., 269, 271, 291, 299, 300; 345 Moroz, T.R., 269; 345 Morozov, N.P., 345; 771 Morris, A.W., 552, 571, 578, 580, 583, 594, 613; 771 See Abdullah, M.I., 593; 743 See Foster, P., 578, 582, 614, 615, 637, 642; 756 Morris, C.P. See Russell, G., 560, 722; 778 Morris, I. See Mague, T.H., 78; 122 Morris, L.J. See Yaro, I., 88; 127 Morris, O.P., 606; 771 Morris, R.J., 76, 78, 88, 608; 123, 771 See Gaskell, S.J., 88; 119 See Joint, I.R., 74–127

Interlinking of physical Morrison, D.R. See Sternberg, R.W., 195 Morrison, S.J. See White, D.C., 127 Morse, A.P. See Komar, P.D., 163, 166; 193 Morse, J.W. See Karl, D.M., 88; 121 Mortensen, J.L., 605; 771 Morton, J.W., 552, 591, 627; 771 Mosher, H.S. See Elam, J.S., 462; 465 Moszynski, A., 226, 231, 283; 346 Mott, J.C. See Pyefinch, K.A., 565; 776 Mountain, C.W. See Gundersen, K., 120 Moyle, P.B. See Grossman, G.D., 457, 458; 467 Mudroch, A., 576, 579; 771 Muehlberger, C., 540, 559; 771 Mueller, D., 543, 545, 552, 718; 772 Muench, K.A. See Thomson, D.A., 437; 471 Mulawka, S.T. See Pringle, B.H., 689; 776 Müller, A., 241; 346 Müller, G., 614, 643; 771 Muller, G. See Förstner, U., 616, 617; 756 Müller, P.J., 84; 123 Muller, R., 428; 469 Munda, I.M., 651, 653, 654; 772 Munday, K.A. See Boulton, A., 566; 748 See Kerkut, G.A., 544, 729; 764 Munier, R. See Craig, Jr, H.L., 752 Munk, O., 489, 526; 532 See Bertelsen, E., 488; 529 See Steenstrup, S., 526; 533 Munro, A.L.S., 94, 334; 123, 346 See Litchfield, G.D., 76; 122 See Steele, J.H., 201, 236; 348 Muntz, W.R.A., 526; 532 Muraoka, J.S., 622; 772 Murdoch, M.B. See Williams, R.B., 715; 787 Murison, D.J. See McIntyre, A.D., 225, 236; 345 Murphy, L.S., 548, 555, 560, 565, 600, 719, 743; 772 Murphy, R.C., 31; 35 Murray, A.P., 641; 772 Murray, C.N., 617; 772 Murray, H.E. See Guthrie, R.K., 556; 760 Murray, J., 138, 411, 516; 194, 417, 532 Murray, J.W., 114; 123 See Bacon, M.P., 190 See Balistreri, L., 185; 190 See Grundmanis, V., 103, 104; 119 Murray, L. See Jones, G.E., 551, 557; 764 See Murray, C.N., 617; 771 Murray, S.N. See Horn, M.H., 457; 467

1336

Oceanography and marine biology

863

Murrish, D.E., 9; 35 Murty, P.S.N., 624; 772 Musil, G. See Hulet, W.H., 489; 531 Mustaffi, Z., 625; 772 Muszynski, G. See Ehrlich, K.F., 446; 465 Mutsaddi, K.B., 438, 442, 457; 469 Muus, B.J., 224, 227, 228, 229, 231, 242, 243, 244, 264, 271, 276, 280, 321, 456; 346, 469 Muzychuk, N.T., 622; 772 Myers, D.N., 542; 772 Myers, G.S. See Greenwood, P.H., 474; 530 Myhal’, O.K., 722; 772 Myhre, K. See Rodsaether, M.C., 777 Myklestad, S., 561, 575, 581; 772 See Melhuus, A., 551; 769 Nafpaktitis, B.G. See McNulty, J.A., 525: 532 See Tsuji, F.I., 533 Naidu, A.S., 606, 642; 772 Naiki, N., 611; 772 Nair, S., 100; 123 Nakahara, M., 713; 772 Nakai, T. See Seki, H., 91; 125 Nakamura, R., 427, 430, 444, 446, 447 453, 460, 461; 469 Nakanishi, M. See Tanaka, N., 97; 126 Nakano, Y. See Cormier, M.J., 494; 529 Nakatani, R.E. See Stober, Q.J., 557; 782 Napora, T.A. See Swift, E., 522; 533 Narasimhamurti, C.C. See Subba Rao, B. V.S.S.R., 348 Narayanan, B. See Ansell, A.D., 369; 415 See Trevallion, A., 356; 417 Naruse, Y., 636; 772 Nathan, A. See Nielsen, S.A., 663, 666 670; 773 National Academy of Sciences, 175; 194 National Marine Fisheries Service Highlands N.J. Mid Atlantic Coastal Fisheries Center, 626, 641, 643, 663, 665, 671, 680, 681; 771 National Oceanic and Atmospheric Administration (NOAA), 539, 591; 772 National Research Council, 556; 772 Naumov, A.G. See Latogurskii, V.I., 30; 35 Navrot, J., 679; 773 See Amiel, A.J., 586, 642; 744 Nealler, E. See Castilla, J.C., 539, 626; 750 Nealson, K. See Reichelt, J.L., 486; 532 Nealson, K.H., 475, 476, 477, 483, 484, 485, 486, 490, 491; 532 See Bang, S.S., 476; 529 See Eberhard, A., 529 See Jensen, M.J., 531 See Leisman, G., 483; 531 See Morin, J.G., 531

Interlinking of physical

1338

Nealson, K.H. See Ruby, E.G., 476, 491; 532 See Tebo, B.M., 485; 533 Nedoclan, G. See Majori, L., 566; 768 Nedwell, D.B., 105; 123 Needham, H.D. See Francheteau, J., 119, 757 Neelakantan, B., 643; 773 Neff, J.M., 564, 729; 773 Neff, J.W., 582, 604, 661, 669, 671, 685, 715; 773 Nehls, H.W. See Arndt, E.A., 242; 338 Neilands, J.B., 583, 608; 773 Nellen, W., 293; 346 Nelson, B.A. See Greig, R.A., 575; 759 Nelson, D.A. See Calabrese, A., 750 See Greig, R.A., 575; 758 Nelson, M.D. See Guarino, C.F., 588; 759 Neshyba, S., 14; 35 See Foster, T.D., 27, 29; 33 Nevissi, A. See Schelle, W.R., 631; 779 Newall, M. See Goodman, C., 555; 759 Newbury, T.K., 588; 773 Newell, P.F., 352, 393, 399, 400, 403; 417 Newell, R.C., 94, 260, 356, 389, 393, 395, 396, 420, 427, 444, 448; 123, 346, 417, 469 See Branch, G.M., 388; 415 Newman, W.A. See Enright, J.T., 115; 119 Newton, A.S. See Anderlini, V.C., 745 Ng, W.-S. See Gordon, M.S., 433, 455; 465 Nicoll, P.A. See Grave, C., 465; 759 Nicholls, G.D., 574, 647; 773 Nichols, B.W. See Yaro, I., 89; 128 Nichols, J.A., 201; 346 Nicol, J.A.C., 473, 499, 501, 503, 509, 522; 532 See Clarke, G.L., 508; 529 See Watson, M., 477, 480; 533 Nickels, J.W. See White, D.C., 127 Nicklès, M., 352; 417 Nickless, G. See Afran, A., 642; 744 See Elderfield, H., 754 See Howard, A.G., 550, 610, 612, 661, 675, 677, 679; 762 See Stenner, R.D., 567, 591, 642, 646, 650, 651, 653, 654, 657, 661, 666, 671, 675, 677, 679, 681, 683, 686, 704; 781 Nielsen, C.O., 234, 237, 239, 269, 276, 317, 334; 346 See Christensen, B., 312; 339 Nielsen, K.N. See Riisgård, H.U., 540; 777 Nielsen, N.B., 581; 773 Nielsen, S.A., 663, 666, 670; 773 Niemi, A.A., 540; 773 Nikolayeva, V.K. See Klenova, M.V., 193 See Vikhrenko, N.M., 138; 197

Oceanography and marine biology Nishikata, J. See Morii, H., 434; 469 Nishikawa, M., 455; 469 Nishikawa, K., 455, 568; 469, 773 See Tabala, K., 567; 783 Nishimura, H. See Hiraizumi, Y., 642; 761 See Kawakami, Y., 643; 764 Nissenbaum, A., 106, 616; 123, 773 See Presley, B.J., 640; 776 Nitsas, K. See Kovatisis, A., 668; 766 Nival, P., 100; 124 Nival, S. See Nival, P., 100; 124 Nixon, S. See Seitzinger, S., 105; 125 Noble, R.G. See Brown, A.C., 352, 365, 399; 416 Noda, H., 538; 773 Noddack, I., 635, 647; 773 Noddack, W. See Noddack, I., 635, 647; 773 Nodot, C. See Bodoy, A., 309; 338 Noel-Lambot, F., 574, 613; 773 Nogami, H. See Kobayashi, N., 557; 765 Nollendorf, A.F. See Pakalne, D.S., 560; 774 See Upitis, V.V., 561; 784 Nolting, R.F. See Duinker, J.C., 594, 613, 617, 637, 641; 754 Nones, N.V. See Jones, L.H., 764 Nonotte, G., 436; 469 Nordstrom, D.K. See Ball, J.W., 604; 746 Norkrans, B., 108, 114; 124 See Dahlbäck, B.O., 108; 118 Norman, E., 622; 773 Normark, W. See Francheteau, J., 119, 757 See Speiss, F.N., 126 Norrell, S.A. See Naidu, A.S., 606; 772 Norris, K.S., 427, 430, 443; 469 North, W.J., 540, 561; 773 See Clendenning, K.A., 539, 561; 751 Northcote, T.G., 693, 696, 700, 702, 703, 706; 773 Norton, A.G. See Bawden, C.A., 591; 746 Norvell, W.A. See Lindsay, W.L., 550; 767 Noskin, V.E., 176, 179; 194 See Bowen, V.T., 191 Novitsky, J.A., 102, 109; 124 See Kepkay, P.E., 102; 121 Nozaki, I.M., 438; 469 Nozaki, Y., 180, 184, 185; 194 See Brewer, P.G., 191 See Spencer, D.W., 84; 126 Nriagu, J.O., 539, 585; 773 Nursall, J.R., 432, 453, 456, 458; 469 Nyberg, D. See Szeto, C., 551; 784

865

Interlinking of physical Nygaard, D.D., 600; 773 O’Brien, C.H., 477; 533 Obukhov, A.I. See Khristoforova, N.K., 659; 765 Ockelmann, K.W. See Berg, K., 233; 338 O’Connor, F.B., 202, 230, 234, 237, 316, 317, 318; 346 See Christensen, B., 217, 312; 339 O’Connor, J.M., 455; 469 O’Day, W.T., 489, 508, 510; 533 Odum, E.P., 87; 124 Oeschger, H. See Williams, P.M., 111; 127 Ogden, S., 607; 773 Oglesby, L.C. See Gordon, M.S., 466 Oglesby, L.S., 313; 346 Ogura, N., 92; 124 Oguri, M. See Soule, D., 562, 565, 626; 781 O’Hara, J., 568; 772 Ohlendorf, H.M., 569; 773 Ohmori, S. See Kusaka, Y., 766 Ohshima, F. See Hosokawa, I., 593; 762 Ohye, R. See Gundersen, K., 120 Okada, K. See Inoue, S., 497; 531 Okaichi, T. See Okutani, K., 641; 774 Okamura, O., 487; 533 Okazaki, R.K., 719; 773 O’Kelley, J.C., 536; 773 Okubo.A., 773; 194 Okutani, K., 641; 774 Olafson, R.W., 550, 567, 574, 582, 585, 613; 774 Oldnall, R.J., 624, 643; 774 Oliff, W.B., 321; 346 Oliver, B.G., 587; 774 Oliver, J.D., 90; 124 Oliver, J.S. See Dayton, P.K., 5; 33 Olley, J. See Ratkowsky, D.A., 663; 776 Olney, C.E. See Duce, R.A., 754 Olofsson, J.A., 600; 774 Olsen, C.R., 774, 643; 774 Olsen, J. See Rodsaether, M.C., 777 Olson, E.A. See Broecker, W.S., 176; 191 Olson, K.R., 563, 718, 733; 774 Olson, R. See Williams, P.M., 88, 106; 127 Olthof, M., 589; 774 O’Neill, T.B., 623; 773 Oostdam, B.L. See Helmy, M.M., 544; 761 Oostenbrink, M., 202; 346 Oota, T. See Dotsu, Y., 437, 440, 442; 465 Ooterhof, D.K. See White, D.C., 127

1340

Oceanography and marine biology Oppenheimer, C.H., 109, 644, 658; 124, 774 Oppenheimer, N.J. See Eberhard, A., 530 Orcutt, J. See Speiss, F.N., 126 Order, O.M., 89; 124 See Johns, R.B., 90; 121 Ordzie, C., 453; 469 Oregioni, B. See Polikarpov, G.G., 565; 776 See Fowler, S.W., 564, 658, 668; 756 Oregioni, D.S. See Polikarpov, G.G., Oren, A., 105; 124 Orliczek, C. See Hopner, T., 593, 609; 762 Orlob, G.T., 604; 774 Oro, J. See Tornabene, T.G., 87, 90; 126 Orren, M.J., 608, 636; 774 See Cuthbert, K.C., 354, 403; 416 See Shannon, L.V., 194 Orton, J.H., 562; 774 Orunesu, M. See Viarengo, A., 786 Oshida, P.S. See Reish, D., 777 Osterberg, C., 580; 774 See Johnson, V., 594; 763 Osteryoung, J., 744; 774 Otobe, H. See Seki, H., 91; 125 Ott, J. See Wieser, W., 274; 349 Otte, G., 321; 346 Ottow, J.C.G., 625; 774 See Heitzer., R.D., 104; 119 Overbeck, J., 99; 124 Overnell, J., 541, 554, 717, 721, 723; 774 Owens, A., 429; 469 Owens, L.D., 469; 774 Owens, T.G. See Christensen, J.P., 111; 118 Oy, E., 651, 654; 774 Ozaki, H., 569; 774 Ozawa, T. See Honma, Y., 438; 467 Ozkan, M.A. See Ramelow, G., 776 Ozoh, P.T.E., 547, 740; 774 Paasivirta, L. See Haka, P., 342 Pace, F., 541, 722; 774 Packard, T.T., 104, 106, 111; 124 See Christensen, J.P., 110, 113; 117 See Codispoti, L.A., 104; 117 See Devol, A.H., 87; 118 See Dugdale, R.C., 118 See Hobbie, J.E., 119 Padan, J.W., 591, 626; 774 Padovani, I.O. de See Lowman, F.G., 767

867

Interlinking of physical Paerl, H.W., 87, 94; 124 Page, A.L. See Sposito, G., 589; 781 Pagenkopf, G.K., 593, 609; 774 Paine, R.T., 458, 460, 462; 469 Pak, H., 105, 136; 124, 194 See Bartz, R., 134; 190 See Carder, K.L., 151; 191 See Kitchen, J.C., 142; 192 Pakalne, D.S., 560; 774 Pakaline, O.S. See Upitis, V.V., 560; 784 Pal, S., 643; 774 Palka, J., 236, 239; 346 Palmer, C.G., 358, 367, 376; 417 Palmer, D.S., 109; 124 Palmer, H.D., 626, 646, 687; 774 Palmer, M.F., 215, 225, 233, 293, 298, 314, 317, 319; 346 Palmer, R.A. See Paine, R.T., 458, 460, 462; 469 Palmer, S.E., 618; 774 Palumbo, A.V., 106; 124 See Ferguson, R.L., 106; 118 Pamatmat, M., 236, 320, 396; 346, 417 Pamatmat, M.M., 113; 124 See Goering, J.J., 104; 119 Pang, P.K.T. See Marusic, E.T., 468 Paoletti, A. See Melluso, G., 615; 770 Papaconstantinou, C., 425, 439; 469 Papadopoulou, C., 663, 664, 668, 672; 774 Parchevskaya, D.S. See Polikarpov, G.G., 565; 776 Parinello, N. See De Leo, G., 440; 465 Parker, B.C. See Cairns, J., 551; 750 Parker, J.H. See Duedall, I.W., 754 Parker, P. See Pulich, W., 579; 776 Parker, P.L., 88, 539; 124, 774 See Goldberg, E.D., 758 See Leo, R.F., 87, 88; 122 Parks, G.A., 618; 774 Parsons, T.R., 91, 141, 550, 613; 124, 194, 775 See Koeller, P., 567, 712, 714, 743; 765 Patchineelam, S.R., 643; 775 See Stoffers, P., 614; 782 Patin, S.A., 540; 775 See Ibragim, A.M., 541; 763 See Morozov, N.P., 763; 770 Paton, D.W. See Thompson, J.A.J., 591, 626, 641; 784 Patouillet, C.E. See Hannan, P.J., 554, 722, 739; 760 Patrick, F.M., 325, 558, 579, 715; 346, 775 Patrick, Jr, W.H., 600, 627; 775 See Khalid, R.A., 615; 764

1342

Oceanography and marine biology Patterson, S.L. See Joyce, T.M., 29; 35 Patterson, T.A. See Leyden, D.E., 767 Pattullo, J. See Osterberg, C., 580; 774 Paskausky, D.F. See Dehlinger, P., 753 Paul, A.C., 753; 775 Paul, J., 641; 775 Paulin, E., 563, 739; 775 Pauly, D., 523; 533 Paus, P.E. See Skei, J., 597, 642; 781 Paxton, J.R. See Graham, P.H., 484; 530 Payne, J.F. See Fletcher, G.L., 538; 756 Payne, J.R. See Stuermer, D.H., 609, 614; 783 Payne, K., 551, 592, 614; 775 Payne, M.R., 7; 35 Peakall, D.B. See Vermeer, K., 648, 650, 657, 666, 684, 715; 786 Peake, E., 609; 775 Pearcy, W. See Osterberg, C., 580; 774 Pearse, J.B., 365; 417 Pearson, T.H., 321, 337; 346 See Miller, D., 76; 123 Pechenik, J.A., 374; 417 Peden, A.E., 422, 423; 469 Peden, J.D., 578; 775 Peddicord, R. See Hirsch, N.D., 591; 761 Peile, A.J., 365; 417 Peirson, D.H., 585, 633; 775 Pellenbarg, R.E., 571, 579, 586, 605; 775 Pempkowiak, J. See Bojanowski, R., 605; 748 Penchaszadeh, P.E., 352, 374, 375; 417 Peng, T.H., 85; 124 See Olsen, C.R., 774 Pennington, C.H. See Wolf, P.L., 576, 579; 788 Penrith, M.-L., 422; 469 Pepelnik, R. See Fanger, H.U., 624; 756 Pequegnat, W.E. See Rezak, R., 777 Percy, J.A., 309, 311; 346 Pereira, J.P. See Paulin, E., 563, 739; 775 Perkins, J. See Leumer, G., 590; 767 Perkins, T.G. See Sposito, G., 610; 782 Perlmutter, S. See Baier, R.E., 108; 117 Permitin, I.U.E., 5; 35 Pern, U. See Laumen, J., 439; 468 Perry, G.J., 89; 124 See Volkman, J.K., 126 Pertica, M. See Viarengo, A., 786 Perur, N.G. See Upadya, G.S., 618; 785 Pervushin, A.S. See Latogurskii, V.I., 30; 35 Pesando, D., 558; 775

869

Interlinking of physical

1344

See Gnassia-Barelli, M., 559; 757 Pesch, C., 551, 614, 630, 671, 724; 775 Peter, R., 775; 775 Peters, T. See Filip, D.S., 580; 756 Petersen, H. See Kremling, K., 617, 638; 766 Peterson, B.J., 99; 124 Peterson, M.H. See Groover, R.E., 590; 759 See Groves, R.E., 622; 758 Peterson, P.J. See Klumpp, D.W., 650, 653, 654, 656, 668, 670, 680, 683; 765 Peterson, R.E., 162; 194 Petkevich, T.A., 578, 684; 775 Petronio, F. See Majori, L., 564, 573; 768 Petrov, Y.M., 635; 775 Pettis, R.W., 624; 775 Petzold, T.J. See Smith, R.C., 134; 195 See Tyler, J.E., 133; 197 Peyton, T., 585; 775 Pfannkuche, O., 201, 210, 212, 214, 215, 220, 221, 222, 223, 225, 226, 228, 229, 232, 235, 236, 237, 244, 245, 247, 253, 254, 257, 258, 261, 263, 265, 268, 269, 271, 277, 278, 279, 280, 281, 282, 287, 288, 289, 290, 291, 292, 293, 295; 346 See Giere, O., 197–350; 223, 263, 268, 269, 288, 289, 291, 300, 301, 303, 325, 336; 341 Pfennig, N. See McInerney, M.J., 105; 123 Pfuderer, H.A. See Braunstein, H.M., 539; 748 Pheiffer, T.H., 642; 775 Phelps, D.K., 577, 714, 716; 775 See Lowman, F.G., 767 See Widdows, J., 556; 787 Philbrick, C.W. See Gurtisen, J.M., 607; 759 Phillip, A.T., 622, 723; 775 Phillips, Jr, A.M., 458; 469 Phillips, C.N.K. See McLusky, D.S., 562, 717, 719; 770 Phillips, D.J.H., 554, 556, 576, 635, 643, 649, 650, 651, 655, 666, 744; 775 Phillips, G.R., 570; 775 Phillips, R.R., 432, 439, 440, 447, 449, 453, 461; 469 Philp, R.P. See Brooks, P.W., 117 Phizacklea, P. See McLusky, D.S., 264, 300; 345 Pickering, W.F., 600; 775 See Farrah, H., 593, 614, 617; 755 See Payne, K., 550, 593, 614; 775 Pierantoni, U., 248; 346 Pierce, R.C. See Spear, P.A., 539; 781 Pierce, Jr, R.H. See Martin, D.F., 543, 610; 769 Pieterse, A., 560; 775 Pietsch, T.W. See Bertelsen, E., 489; 529 Pienkowski, M.W. See Evans, P.R., 244; 341 Piezzi, R.S., 31; 35 Pilkey, O.H., 574; 775 Pillai, K.C. See Paul, A.C., 775; 775

Oceanography and marine biology Pilotte, J.O., 643; 776 Pilson, M.E.Q. See Betzer, S.B., 675, 689; 747 See Seitzinger, S., 104; 125 Piltz, F. See Reish, D.J., 547; 777 Piltz, F.M. See Martin, M.J., 544; 769 See Reish, D.J., 586; 777 Pinchuk, V.I., 422, 423, 443; 469 Pingree, R.D., 58, 543; 73, 776 Piotrowicz, S.R., 586, 597, 636, 639; 776 See Duce, R.A., 576, 602; 753 Piper, D.Z., 643; 776 Pirie, B.J.S. See George, S.G., 758 Pirt, S.J., 101; 124 Pitt, Jr, W.W. See Katz, S., 611; 764 Piuze, J. See Cossa, D., 593; 752 Plagmann, J., 243; 346 Platt, T. See Nealson, K.H., 476; 532 Pocklington, R., 602; 776 Poddubnaya, T.L., 209, 212, 215, 221, 238; 346 Podgorskaya, E.D. See Kazaikin, N.I., 586; 765 Pointier, J.P. See Berland, B.R., 554; 747 Pojasek, R.B., 593; 776 Polikarpov, G.G., 565; 776 Pollingher, U., 94; 124 Polloni, P.T. See Rowe, G.T., 113; 125 Polo, B. See Barbare, A., 746 Pomeroy, L.R. See Hobbie, J.E., 120 See Wiebe, W.J., 86, 87; 126 Pomroy, A.J. See Stebbing, A.R.D., 555, 726; 782 Pope, D. See Smith, W., 109; 126 Pope, D.H., 109; 124 See Landau, J.V., 110; 122 Pope, D.M. See Fink, Jr, L.K., 122; 756 Popham, J.D. See Stump, I.G., 783 Port, J. See Timoney, J.F., 547, 644; 785 Porter, K.G., 86; 124 Portmann, J.E., 591, 626, 691; 776 Pörtner, H.-O., 320; 347 Porumb, I.I., 457; 469 Postgate, J.R., 101; 124 Postma, H. See Manuels, M.W., 88; 122 Potts, G.W., 421, 422, 423; 469 Potts, W.T.W., 410; 417 Poulet, S.A. See Cossa, D., 594; 752 Pouliot, D. See Cossa, D., 570; 752 Poulsen, E.M., 493; 533 Powell, E.N., 281; 347 Prakash, A., 543, 609, 715; 776

871

Interlinking of physical See Rashid, M.A., 609; 776 See Sheldon, R.W., 130; 194 Prejs, A. See Kajak, Z., 201; 343 Preslan, J.E. See Hamilton, R.D., 91; 120 Presley, B.J., 640, 641; 776 See Brooks, R.R., 776; 749 See Nissenbaum, A., 106; 123 See Sims, Jr, R.R., 679, 681, 689, 691, 692; 780 See Trefry, J.H., 591, 626; 784 Price, Jr, K.S. See Derickson, W.K., 421, 422; 464 Price, M. See Montgomery, J.R., 577; 771 Price, M.T. See Montgomery, J.R., 577, 587, 715; 771 Price, N.B. See Duchart, P., 771; 754 See Skei, J.M., 587; 780 See Sholkovitz, E.R., 593; 780 Price, R.K.J., 546; 776 Pringle, B.H., 662, 663, 665, 689; 776 See Shuster, C., 776; 780 Pritchard, P.H. See Jannasch, H.W., 87; 121 Prosi, F., 536; 776 Provasoli, L., 538, 578; 776 See Hutner, S.H., 537; 762 Pruefer, P. See Auerbach, S., 554; 745 Prytherch, H.F., 538, 636, 733, 742; 776 Pugh, P.R., 141, 151, 152; 194 See Pingree, R.D., 542; 775 Pugh-Thomas, M. See Eyres, J.P., 575; 756 Pugno, P. See Gross, R.E., 540; 759 Pulich, W., 579, 648, 657; 776 Purushan, K.S. See Sankranarayanan, V. N., 562; 778 Pyefinch, K.A., 565; 776 Pyzik, A.J. See Sommer, S.E., 642, 643; 781 Qasim, S.Z., 441; 469 Quigley, M.M., 114; 125 Quinn, J.G. See Duce, R.A., 754 See Kerr, R.A., 601, 743; 764 See Meyers, P.A., 613; 770 See Schultz, D.M., 88; 125 See Van Vleet, E.S., 88; 126 Raa, J. See Rodsaether, M.C., 777 Rabe, F.W. See Bartlett, L., 619; 746 Rabotnova, I.L. See Avakyan, Z.A., 559, 721; 745 Rachlin, J.W. See Rosko, J.J., 560, 717, 721; 778 Radford, P.J. See Warwick, R.M., 233, 240; 349 Radkewitsch, A., 253; 347 Radlett, A.J. See Jones, L.H., 764

1346

Oceanography and marine biology Radwin, G.E., 376; 417 Radziejewska, T., 336; 347 Ragotzkie, R.A. See Keillor, J.P., 591, 628; 765 Ragsdale, H.L., 592; 776 Rainbow, P.S., 570, 574; 776 Rainer, S.F., 408; 417 Rains, D.W. See Ruddell, C.L., 546, 549, 585, 644; 778 Rajendran, A., 540, 717, 721; 776 Rakusa-Suszczewski, S., 9; 35 Rama, S. See Bhat, S.G., 191 Ramachandran Nair, P.V., 422, 428; 470 Ramamoorthi, K. See Kumaragura, A.K., 562, 732; 766 Ramamoorthy, S. See Briand, F., 611; 749 See Laube, V., 603; 766 See Manning, P.G., 608; 768 Ramelow, G., 681, 689, 691, 692; 776 Ramm, A.E., 105; 125 Ramnarine, A. See Lewis, A.G., 556, 562, 578; 767 Ramondetta, P.J., 642; 776 Ramsay, J.A., 313; 347 Ramsay, A.J., 86, 90; 124 Ramsay, B.A. See Sprague, J.B., 568, 737; 782 Rand, G. See Kleerekoper, H., 567; 765 Randall, C.W., 588; 776 Randall, J.E. See Hori, K., 467 Rangin, C. See Francheteau, J., 119, 757 See Speiss, F.N., 126 Rankin, P.C. See Glasby, G.P., 624; 758 Rao, C.M. See Murty, P.S.N., 624; 772 Rao, T.S.S. See Sankranarayanan, V.N., 562; 778 Rashid, A. See Prakash, A., 608, 715; 776 Rashid, M.A., 608, 609; 776 Ratkowsky, D.A., 663; 776 Rau, G.H., 115; 125 Rauschenplatt, F.M., 242; 347 Rauser, W.E. See Hogan, G.D., 548; 761 Rauther, M., 515; 533 Ravera, O., 281; 347 Ray, B.J. See Duce, R.A., 585; 754 Ray, B.Y. See Piotrowicz, S.R., 775 Ray, D.L., 509; 532 Ray, M.J. See Reish, D.J., 777 Raymond, J.A. See Wright, W.G., 427, 436; 471 Raymont, J.E.G., 539; 777 Rayner, S. See Tamblyn, N., 622; 784 Reddy, C.A. See Bryant, M.P., 105; 118 Reddy, C.V.G. See Murty, P.S.N., 624; 772 See Sankranarayanan, V.N., 593, 635; 778

873

Interlinking of physical

1348

See Zingde, M.D., 789 Redfield, H. C, 537; 777 Redmond, J.R., 537; 777 Reeburgh, W.S., 106; 125 Rees, C.B., 243; 347 Rees, C.P. See Collinson, R.I., 564, 643; 751 Reeve, M.R., 544, 552, 556; 777 Rehm, E. See Halbach, P., 624; 760 Reichelt, J.L., 475, 476, 486, 490, 491; 533 Reid, S.M. See Bentley-Mowat, J.A., 541, 560; 747 Reimchen, T.E., 461; 470 Reimer, L.W., 409; 418 Reineck, H.-E. See Dominik, J., 588; 753 See Förstner, U., 642; 756 Reinhart, W. See Sorem, R., 625; 781 Reise, K., 241, 242, 243, 282, 336, 461; 347, 470 See Scherer, B., 243; 347 Reish, D.J., 539, 544, 547, 562, 587, 600, 644, 647, 649, 651, 660, 675, 681, 687, 689, 721, 723, 724; 777 See Bellan, G., 327; 338 See Bellan-Santini, D., 729; 747 See Martin, M.J., 544; 768 Remane, A., 274; 347 Renaud-Debyser, J., 267; 347 Renaud-Mornant, J. See Lassèrre, P., 280, 313, 320; 344 Rendell, P.S., 615; 777 Renger, E.H. See Harrison, W.G., 540; 761 Revelante, N. See Gilmartin, M., 697, 707; 758 Revsbech, N.P. See Sørensen, J., 105; 126 Reynolds, B. See Pesch, C., 775 Reynolds, B.H., 659, 661, 671; 777 Reynolds, C.S., 154; 194 Reynolds, G.T. See Case, J.F., 529 Reynolds, W.R., 643; 777 Reynoldson, T.B., 203, 217, 218, 219, 245, 247, 253, 254, 269, 297, 300, 323, 324, See Sigurjonsdottir, H., 245; 348 Rezak, R., 672, 699, 707; 777 Rhee, G.Y., 96; 125 Rheinberger, R., 584; 777 Rheinheimer, G., 237; 347 Rho, J. See Berger, P.S., 611; 747 Rhoads, D.C., 281; 347 See Yingst, J.Y., 236, 279, 281; 350 Rice, Jr, D.W., 547, 567, 735; 777 See Harrison, F.L., 546, 576, 580, 582, 658, 717, 719, 731, 738; 760 Rice, E.L. See Coleman, R.D., 570; 751 Rice, J.C. See Brazda, F.G., 318; 338 Rice, T.R. See Wolfe, D.A., 570; 788

Oceanography and marine biology Richards, F.A., 104, 601; 124, 777 See Codispoti, L.A., 106; 117 See Packard, T.T., 106, 110; 124 Richards, K.S., 291; 347 See Fleming, T.P., 324; 341 Richardson, M.D., 5; 35 Richardson, M.J., 136, 141, 142; 194 See Dymond, J., 191 Richez, C. See Lacombe, H., 57; 73 Richkus, W.A., 432, 446, 449, 453; 470 Richter, K., 313; 347 Richter, R.D. See Theis, T.L., 590; 784 Ricketts, E.F., 420; 470 Ridley, M., 441; 470 Riedl, R.J. See Fenchel, T., 105; 119 Rieger, R.M. See Powell, E.N., 281; 347 Riegle, K.C., 433, 435; 470 See Horn, M.H., 427, 430, 432, 433, 435, 442; 466 Riemann, F., 240; 347 Rieper, M., 100; 125 Righi, G., 206; 347 Riisgård, H.U., 540; 777 Riley, G.A., 594, 636; 777 Riley, J.D., 422, 449; 470 Riley, J.P., 602, 637, 645, 660, 687, 688; 777 See Mantoura, R.F.C., 582, 602; 768 See Segar, D.A., 564; 780 Rippey, E. See Kilburn, R.N., 351, 358; 417 Risebrough, R.W. See Anderlini, V.C., 569; 745 See Goldberg, E.D., 758 See Ohlendorf, H.M., 568; 772 Ritchie, L.S. See Frick, L.P., 563; 758 Ritter, C.J. See Bergman, S. C., 593; 749 Ritter, R.B., 622; 777 Robb, Jr, A.E., 715; 777 Roberts, B.L. See Denton, E.J., 519; 529 Robertson, K.J. See Williams, P.M., 588; 784 Robertson, W. See Goldberg, E.D., 759 Roche, P.M. See Jones, G.E., 550, 720; 764 Rochford, D.J. See Wyrtki, K., 9, 14; 36 Rodgers, J. H., 561; 777 Rodsaether, M.C., 546; 777 Roegge, M.A., 622; 778 Roesijadi, G., 574; 778 Rofritz, D.J., 244; 347 Rogerson, P. See Pesch, C., 775 Roman, M.R. See Honjo, S., 82, 142, 167; 121, 193 Romanenko, V.I., 99; 125

875

Interlinking of physical Romanov, A.S., 637; 778 Romeo, M. See Gnassia-Barelli, M., 560; 758 Romeril, M.G., 553, 589, 635, 654; 778 See Boyden, C.R., 638, 642, 653, 661, 677, 683, 684; 748 Romesser, J.A., 106; 125 Roper, C.F.E. See Young, R.E., 523, 525; 534 Roper, D.S., 455; 470 Ropes, J.W. See Wenzloff, D.R., 571; 787 Rose, R.R. See Minear, R.A., 600; 771 Rose, W.C., 536, 647; 778 Rosen, A.A. See Katz, S., 611; 764 Rosen, D.E. See Greenwood, P.H., 474; 530 Rosen, W., 613, 618; 778 See Williams, P.M., 608; 787 Rosenbauer, R.J. See Bischoff, J.L., 625; 747 Rosenberg, R., 627; 778 See Pearson, T.H., 320, 322, 337; 345 Rosenblatt, R.H., 480, 481, 483; 533 See Graham, J.B., 435; 465 Rosenfeld, J.K. See Martens, C.S., 106; 123 Rosin, D. See Stolzberg, R.J., 606; 783 Roskam, R.T., 585; 778 Rosko, J.J., 560, 717, 722; 778 Ross, E.H. See Deuser, W.G., 154; 192 Rossi, S.S., 330; 347 See Reish, D., 777 Rossell, R.M. See Eisler, R., 538; 754 Rossouw, G. See McLachlan, A., 352; 417 Roth, I., 635, 638, 643, 697, 701, 708, 710, 712; 777 See Riley, J.P., 645; 777 Rothschild, Lord, See Barnes, H., 635; 746 Rowe, F. See Smith, D.F., 125 Rowe, G.T., 114, 154, 157; 125, 194 See Smith, K.L., 125 See Staresinic, N., 158; 194 Royals, H.E. See Hestand, R.S., 542; 761 Royle, L.G. See Abdullah, M.I., 593, 615 617; 744 See Jones, G.E., 551, 558, 644; 764 Rozhanskaya, L.I., 645, 666, 682, 714778 Rublee, P. See Ferguson, R.L., 106; 119 Ruby, E.G., 477, 487, 488, 490, 491; 533 Ruck, J.G., 440, 441; 470 Rucker, J.B., 574; 778 Rudd, J.W.M., 105; 125 Ruddell, C.L., 546, 549, 585, 644; 778 Rueter, J.G. See Morel, N.M.L., 538, 541; 771 Ruivo, M., 539; 778 Rukhadze, Ye. G. See Kobot’yev, V.G., 559; 765

1350

Oceanography and marine biology Rumsby, M.G. See Brooks, R.R., 579, 581, 666, 670; 749 Russell, G., 555, 560, 722; 778 See Fielding, A.H., 542; 755 See Goodman, C., 554; 758 See Morris, O.P., 605; 770 Russell, P. See Klein, D.H., 590; 765 Russell-Hunter, M. See Russell-Hunter, W. D., 381, 382; 418 Russell-Hunter, W.D., 381, 382; 418 See Browne, R.A., 376; 415 Russo, R.C. See Phillips, G.R., 570; 775 Ruthven, J.A., 561, 715; 778 Rutkovskiy, V.M., 624; 778 Rutledge, W.P. See Roegge, M.A., 622; 778 Ruyter van Steveninck, E. de, 235; 346 Ryabinin, A.I. See Romanov, A.S., 637; 778 Rydell, H.S. See Bonatti, E., 614; 748 Ryst, P. van der See McLachlan, A., 201; 344 Rystad, B. See Jensen, A., 548, 720, 722, 723; 764 Ryther, J.H., 98, 542; 125, 778 See Barber, R.T., 578, 601, 605; 747 See Menzel, D.W., 110; 123 See Sanders, J.G., 542; 778 Ryu, B.S., 424; 470 Saarista, P. See Haka, P., 342 Sabater, A.R. See Alted, M.D., 422; 463 Sabharwal, P.S. See Pieterse, A., 560; 775 Sachs, P.L. See Brewer, P.G., 191 See Honjo, S., 82, 156, 168; 119, 192 See Meade, R.H., 123 See Spencer, D.W., 160, 585; 194, 781 Sackett, W.M., 130, 134, 136, 175, 179; 194 See Degens, E.T., 84; 118 See Feely, R.A., 82; 118 Saenko, G.N., 581; 778 Saffé, F. See Theede, H., 301; 349 Sagawa, T. See Kusaka, Y., 635; 766 Saifullah, S.M., 541, 559; 778 Saiga, Y. See Johnson, F.H., 531 Saijo, Y., 20; 35 Saiki, M. See Nakahara, M., 713; 772 Saino, T., 84; 125 Sainsbury, M. See Hardisty, M.W., 713; 760 Sakamoto, S., 538; 778 Saksena, V.P., 437; 470 Salanki, J., 543; 778 Sale, P.F., 422, 423; 470 Saliba, L.J., 549, 564, 720, 739; 778

877

Interlinking of physical

1352

Saliot, A. See Barbier, M., 107; 117 See Marty, J.C., 122 Sallman, B. See Gerchakov, S.M., 590, 622; 758 Salo.A., 601; 778 Salomons, W., 595, 641, 643; 778 Samain, J.F. See Alayse-Danet, A.M., 539; 744 Samylin, A.F., 568, 737; 778 Sanchez, I., 564; 778 Sandberg, J. See Curl, H., 111; 119 Sanders, H.L., 260, 336; 347 Sanders, J.G., 542; 778 See Ryther, J.H., 542; 778 Sanderson, P.L. See Goulder, R., 554; 759 Sankaranarayanan, V.N., 562, 579, 592, 636, 663, 689; 778 See Ansell, A.D., 415 Santiago, R.J. See Montgomery, J.R., 593, 605, 614, 636; 771 See Phelps, D.K., 576; 775 Sargent, J.R. See Cowey, C.B., 458; 465 Sarin, M.M. See Krishnaswami, S., 82 185; 122, 193 Sarker, A.L., 457; 470 Sarmiento, J.L., 180; 195 Sarro, F. See Castagna, A., 668; 750 Sarvala, J., 243; 347 See Haka, P., 342 See Holopainen, I.J., 201; 342 Sarvala, M. See Haka, P., 342 Sasaki, T., 424, 443, 450, 460; 470 Sasaki, Y. See Matsuike, K., 24; 35 Satake, M. See Matsumoto, T., 769 Sato, M. See Yamamoto, Y., 575; 788 Saunders, R.L., 567; 778 See Sprague, J.B., 567; 781 Saward, D., 713; 778 Sawyer, P.J., 540, 559; 778 Saxen, R. See Salo, A., 601; 778 Saydam, C. See Ramelow, G., 776 Schack, R.P. See Leumer, G., 590; 767 Schafer, H.A., 587; 779 See Hershelman, G.P., 586; 762 Schatz, A. See Hutner, S.H., 538; 763 Schaudinn, J. See Theede, H., 301; 349 Schaumberg, G.D. See Sposito, G., 610 782 Scheinberg, I.H., 612; 779 Schelle, W.R., 631; 779 Scheltema, R.S., 375; 418 Schen, J.C., 109; 125 Schenck, P.A. Set-Boon, J.J., 89; 117 Scherer, B., 243; 347

Oceanography and marine biology

879

Scherfig, J. See Christensen, E.R., 542, 588, 631; 751 Schick, L.L. See Fink, Jr, L.K., 751; 756 Scheimer, F. See Wieser, W., 274; 349 Schipp, R., 573, 674; 779 Schlesinger, R.B., 571, 594; 779 Schleyer, M.H., 97; 125 Schmidt, D.J., 554; 779 Schmidt, R., 296; 347 Schmidt, R.L., 536, 539, 574, 579, 591, 600, 602, 612, 613, 617, 629, 631, 634, 636, 641, 642, 643, 648; 779 Schmidt-Moser, R., 243; 347 Schmidt-Nielsen, B. See Ferraris, J.D., 314; 341 Schmidt-Nielsen, K., 391, 393; 418 Schneider, E. See Goldberg, E.D., 758 Schneider, J., 559, 718, 721; 779 Schnier, C., 625, 640; 779 Schnitzer, E., 605; 779 Schoffeniels, E., 537; 779 See Boone, W.R., 537; 748 Scholten-Koerselman, I.J. See Zevenhuizen, L.P.T.M., 550; 789 Schöne, C., 203, 217, 236, 239, 297, 305, 307, 312; 346 Schottel, J. See Silver, S., 550; 780 Schöttler, U., 320; 346 Schrader, H.-J., 81, 142, 167; 125 Schrage, M. See Riemann, F., 240; 347 Schram, M. See McLachlan, A., 352; 417 Schramel, J.R. See Gross, M.G., 759 Schreck, C.B., 545, 557; 779 Schroeder, P.B. See Thorhaug, A., 590; 784 Schroff, G. See Schöttler, U., 319; 346 Schubel, J.K., 637; 779 Schubel, J.R. See Burrell, D.C., 570, 579, 658; 750 See Swift, D.J.P., 137; 194 Schubert, H.R. See Carlucci, A.F., 104; 118 Schubert, J., 581, 619; 779 Schulte, E., 439; 470 Schultz, D.M., 88; 125 Schultz, W., 263, 266, 267, 268, 282, 287, 297, 305; 346 Schumacher, A., 348 Schuster, I., 583; 779 Schuster, R., 245; 348 Schwarz, F.J. See Kirchgessner, M., 619; 765 Schwarz, J.R., 112, 114; 125 See Swartz, R.W., 110; 126 Schwarzer, R.R. See Wheeler, R.B., 588; 787 Schwassman, H.O., 455; 470 Science Applications, Inc., La Jolla, Ca., 654, 659, 668, 675, 683, 686, 688, 742; 779 Sclater, F.R. See Boyle, F.A., 585; 748

Interlinking of physical Scott, A.G. See Rainbow, P.S., 570, 574; 776 Scott, C.D. See Katz, S., 611; 764 Scott, D.M., 563, 734; 780 Scott, M.A. See Howmiller, R.P., 323; 343 Scott, W.D. See Cattell, F.C.R., 585, 634; 750 Scott Van Dyke, J. See Winner, R.W., 557; 787 Scrudato, R.J., 584; 780 Seamark, R.F. See Griffiths, D., 31; 34 Seehaus, H.M. See Bergman, H., 593; 747 Seeliger, U., 561, 610; 780 Sefton, A.D. See Reynoldson, T.B., 245; 347 Segal, E., 275; 348 Segar, D.A., 563, 661, 664, 666, 670, 675, 678; 780 See Riley, J.P., 659, 686, 688; 777 Séguret, M. See Francheteau, J., 119, 757 Seibert, D. See Thomas, W.H., 541; 784 Seibert, D.L.R. See Thomas, W.H., 560; 784 Seifert,R., 263; 348 Seip, K.L., 551, 576; 780 See Melhuus, A., 551; 769 Seitzinger, S., 105; 125 Seki, H., 91, 106, 111; 125 See Parsons, T.R., 142; 194 Seligman, P.F. See Zirino, A., 637; 789 Seng, T.N. See Whitley, L.S., 237; 350 Serne, R.J., 595, 626, 628, 636, 641, 644; 780 Settlemyre, J.L. See Gardner, L.R., 588; 757 Sevenhuysen, W. See Duursma, E.K., 602; 754 Severy, H.W., 636; 780 Shackleton, L.R.E. See McCance, R.A., 675, 677, 679; 769 Shackleton, N. See Peng, T.H., 85; 124 Shackley, S.E., 442; 470 Shakuntala, K., 537; 780 Shand, I.G. See Davis, J.C., 557; 752 Shanks, A.L. See Silver, M.W., 84; 126 Shannon, L.V., 178; 195 Sharp, J.H., 77; 125 Shatila, T.A. See Shiber, J.G., 650, 652, 653, 661, 678; 780 Shatzman, A.R., 537; 780 Shaughnessy, D. See Staresinic, N., 159; 195 Shaw, N., 89; 125 Shear, H. See Hodson, P.V., 539; 761 Sheldon, R.W., 85, 131, 141, 160; 125, 195 See Hobbie, J.E., 119 See Swift, D.J.P., 137; 194 Shen, J. See Gundersen, K., 120 Sheppard, C.R.C., 563, 687; 780 Sherberger, F.F. See Cherry, D.S., 577; 751

1354

Oceanography and marine biology

881

Sherwood, M.J., 545; 780 See Alexander, G.V., 744 Shewan, J.M. See Hendrie, M.S., 475; 531 See Hodgkiss, W., 85; 119 Shiber, J.G., 549, 636, 650, 652, 653, 661, 678, 685, 688; 780 Shields, J. See Raymont, J.E.G., 539; 777 Shilo, M., 477; 533 See Yetinson, T., 476; 533 Shimizu, M. See Taguchi, M., 712; 784 Shimomura, O., 495, 497; 533 See Dunlap, J.C., 497; 529 See Haneda, Y., 493; 530 See Johnson, F.H., 531 Shimp, S.L. See Carlucci, A.F., 87; 118 Shiogaki, M., 437, 440, 442; 470 Shiotsuki, K. See Arai, R., 426; 464 Shipman, W.H. See Dempster, R.P., 622; 753 Shirokova, E.L. See Telitchenko, M.M., 540; 784 Shiroyama, T. See Miller, W.E., 555; 771 Shokes, R. See Callahan, R.A., 653, 655, 660, 664, 668, 673, 675, 677, 680, 684, 686, 688; 750 Sholes, T.S. See Beer, R.M., 160; 191 Sholkovitz, E.R., 593; 780 Shores, D.L. See Backus, R.H., 528 Shrestha, K. See Morales, E., 685; 771 Shroy, R.E. See Duedall, I.W., 754 Shuman, M.S., 754; 780 Shumway, S.E., 395; 418 See Crisp, M., 392; 416 Shurova, N.M., 206; 348 Shuster, C., 348; 780 Shuto, T., 376; 418 Sibert, J., 91, 93; 125 Sibert, J.R. See Brown, T.J., 100; 118 Sick, L.V., 551, 597, 637; 780 Sidner, B.R. See Bouma, A.H., 627; 748 Sie, E.H.-C., 493; 533 See Haneda, Y., 492; 530 See Johnson, F.H., 493; 531 Siebers, D., 240, 325, 544; 348, 780 Sieburth, J.McN., 86, 87, 98, 99, 101, 106, 107, 108, 112, 602, 611; 780 See Burney, C.M., 76; 117 See Hinga, K.R., 192 Siedler, G., 45, 57, 64; 73 Siegal, A., 583, 715; 780 Siegfried, W.R. See Marsh, B., 422, 445, 461; 468 Siegrist, H.G. See Sommer, S.E., 588; 781 Sierra, B., 425; 470 Siesser, W.G., 624; 780

Interlinking of physical Sievers, H.A., 29; 35 Sigurd, M. See Haug, A., 554; 761 Sigurjonsdottir, H., 245; 348 Silker, W.B. See Bishop, J.K.B., 117, 191 Silver, M.W., 84; 126 Silver, S., 549; 780 Silverberg, N. See Lambert, C.E., 194 Sim, R.G. See Olafson, R.W., 574; 774 Simkina, R.G. See Turpaeva, E.P., 549; 785 Simmons, D.R. See Mathews, T.D., 659; 769 Simoneit, B.R., 609; 780 Simonov, A.I. See Turekian, K.K., 539; 785 Simons, R. See Day, J.H., 406; 416 Simpson, C.F. See Cardeilhac, P.T., 545, 719; 750 Simpson, H.J., 586, 642; 781 See Olsen, C.R., 774 Simpson, W.R., 129–95 Sims, Jr, R.R., 681, 690, 691, 692; 781 See Trefry, J.H., 626; 784 Singbal, S.Y.S. See Zingde, M.D., 789 Singer, P.C., 604, 606; 781 Singleton, F.L., 548, 558; 781 Sin’kov, N.A. See Saenko, G.N., 580; 778 Sittig, M., 743; 781 Sivadas, P. See Ansell, A.D., 369; 415 See Trevallion, A., 356; 417 Sivalingam, P.M., 554, 649, 650, 651, 653, 656; 781 Sizemore, R.K. See O’Brien, C.H., 477; 533 Skei, J., 597, 642; 781 Skei, J.M., 588; 781 Sladen, W.J.L., 7; 35 Slawyk, G., 22, 26; 35 Sleep, J.A. See Davis, A.G., 540; 752 Slinn, W.G.N., 781 Sloan, H.D. See Bender, M.E., 570; 747 See Hugget, R.J., 578; 762 Slowey, J.F., 582; 781 See Neff, J.W., 582; 772 Small, L.C. See Cutshall, N., 594; 752 Small, L.F., 166, 580; 195, 781 See Fowler, S.W., 81, 165, 168; 118, 192 See Komar, P.D., 163, 165; 193 Smayda, T.J., 81, 165, 166; 126, 195 Smidt, E.L.B., 242; 348 Smith, A.L., 408; 418 Smith, B.N., 579; 781 Smith, C.A. See Carpenter, J.M., 554, 739; 750 Smith, C.L. See Brewer, P.G., 191

1356

Oceanography and marine biology Smith, D.F., 100; 126 See Wiebe, W.J., 97; 126 Smith, E.L. See White, A., 537; 787 Smith, G.C. See MacCarthy, P., 603; 768 Smith, J.B. See Bader, R.G., 615; 746 Smith, J.D. See Blutstein, H., 604, 636, 637; 748 Smith, J.E. See Jennett, J.C., 560; 764 Smith, K.L., 113; 126 Smith, L.W. See Gardener, W.S., 333; 341 Smith, M.J. See Manahan, S.E., 537, 603; 768 Smith, P.C., 58, 62; 73 Smith, R. See Windom, H., 787 Smith, R.C., 134; 195 Smith, Jr, R.G., 607; 781 See Windom, H.L., 635; 787 Smith, R.G. See Windom, H.L., 574, 575, 663; 787 Smith, R.L., 432, 440, 442, 457; 470 See Barber, R.T., 604, 605; 747 See Dugdale, R.C., 118 Smith, R.N. See Gross, M.G., 759 Smith, S. See Sternberg, R.W., 134; 195 Smith, W., 109; 126 Smith, W.O., 97; 126 Smith, W.P. See Pope, D.H., 109; 124 Smith-Vaniz, W.F. See Colin, P.L., 480; 529 Smokler, P.E. See Young, D.R., 590; 789 Smol, N. See Heip, C., 201; 342 Snodgrass, J.M. See Boden, B.P., 518; 529 So, C.L., 538, 587; 780 Søgaard-Jensen, B. See Riisgård, H.U., 541; 777 Solbé, J.F. de L.G. See Williams, N.V., 203, 217, 317; 349 Solov’eva, L. See Babinets, A.E., 746 Somayajulu, B.L.K., 176, 177; 194 See Craig, H., 180; 191 See Krishnaswami, S., 142, 185; 193 See Lal, D., 176; 193 Somiya, H., 488, 526; 533 Sommer, S.E., 588, 642, 643; 781 Sonis, S. See Jackim, E., 568; 763 Soper, A.E. See Eisler, R., 755 Sopott, B., 245; 348 Sorem, R., 625; 781 Sørensen, J., 105; 126 Sørensson, F. See Norkrans, B., 108; 124 Sorokin, Y.I., 98, 99, 100, 103; 126 Sosnowski, S.L., 552, 556, 565, 728, 744; 781 Soule, D., 562, 565, 626; 781 Soulsby, P.G., 587; 781

883

Interlinking of physical

1358

Soundaraj, R. See Sreenivasan, A., 567; 781 Soutar, A., 155, 159, 166; 195 See Bruland, K.W., 158; 191 See Dymond, J., 191 See Krishnaswami, S., 193 Southwell, C.R., 622; 781 Sovga, E.E. See Babinets, A.E., 746 Spanier, J. See Timoney, J.F., 548; 785 Sparks, A.K. See Martin, S.G., 563; 769 Sparrow, B.W. See Corner, E.D.S., 564; 751 Spaul, E.A. See Palka, J., 236, 239; 346 Spear, P.A., 539; 781 Spears, L.G., 622; 781 Speiss, F.N., 115; 126 Spencer, D.W., 83, 160, 176, 177, 182, 185, 189, 560, 585, 637; 126, 195, 781 See Bacon, M.P., 185; 190, 191 See Brewer, P.G., 191 See Gardner, W.D., 192 Spender, D. See Meade, R.H., 123 Sperber, C., 210; 348 Speyer, M.R. See Hummel, B.L., 491; 763 Spiedo, H. See Bienati, N.L., 26; 33 Spies, R. See Milanovich, F.P., 658; 770 Spies, R.B., 333; 348 See Davis, P.H., 333; 340 Spiess, E. See Romesser, J.A., 125 Spinrad, R.W. See Zaneveld, R.J.V., 136; 195 Spittler, P. See Eick, K., 204; 341 Sponder, D.L., 459; 470 Sposito, G., 589, 610; 781 See Baham, J., 609; 747 Sprague, J.B., 551, 567, 737; 782 See Black, G.A.P., 558; 747 See Howarth, R.S., 551; 762 See Saunders, R.L., 567; 778 Springett, J.A., 204, 217, 224; 348 Spronk, N., 563; 782 Sreenivasan, A., 568; 782 Stahl, M.S., 454; 470 Stam, A. See De Wolf, P., 536; 753 Stanbury, F.A. See Barnes, H., 565; 746 Standen, V., 202, 203, 217, 224, 234; 348 Stankowska-Radziun, M. See Radziejewska, T., 336; 347 Stanley, S.O. See Miller, D., 76; 123 Staresinic, N., 159; 195 Starinkova, O.B. See Petrov, Y.M., 775 Starostin, I.V., 622; 782 Starr, T.J., 722; 782

Oceanography and marine biology

885

Statham, P.J. See Harris, J.E., 556; 760 Staude, C.P. See Armstrong, J.W., 422; 464 Stebbing, A.R.D., 548, 555, 726, 739; 782 Steele, D.H. See Wells, B., 459; 471 Steele, J.H., 78, 79, 201, 236; 126, 348 See Gamble, J.C., 743; 756 Steemann Nielsen, E., 90, 540, 560, 578, 603, 715, 720, 721; 126, 782 Steen, J.B. See Rodsaether, M.C., 546; 777 Steenstrup, S., 526; 533 Stehn, B.D. See Norkrans, B., 114; 124 Steinberg, M.A., 539; 782 Steinhart, H. See Kirchgessner, M., 619; 765 Stenhouse, M.C. See Williams, P.M., 84; 127 Stenner, R.D., 567, 592, 642, 647, 650, 651, 653, 654, 658, 662, 666, 673, 677, 679, 682, 683, 687, 705; 782 See Elderfield, H., 754 Stenström, T.A. See Kjelleberg, S., 108; 122 Stepanyuk, I.A. See Petkevich, T.A., 578, 684; 775 Stephens, Jr, J.S., 446, 447, 451, 452; 470 Stephenson, J., 197, 198, 207, 219, 236, 238, 247, 248, 252, 254, 267, 275; 348 Stephenson, M.D. See Martin, M., 553; 769 AA Stephenson, R.L., 659, 675; 782 Stephenson, R.R., 550, 616; 782 Stepien, C.A. See Demski, L.S., 503; 529 Stern, D., 410, 592, 614, 626; 418, 782 Stern, E. M, 410; 418 Sternberg, R.W., 134; 195 Sternshein, D. See Bovee, E.C., 718; 748 Stevens, D.G., 546, 551; 782 Stevens, J.D., 693, 695, 699, 706, 708, 711; 782 Stevenson, L.H. See Erkenbrecker, C.W., 106; 119 Stewart, G.L. See Beers, J.R., 558; 746 Stickney, R. See Windom, H., 787 Stickney, R.R., 647, 682, 684, 686, 694, 696, 697, 700, 709, 711; 782 Stirling, A. See Saward, D., 713; 778 See Ansell, A.D., 369; 415 Stirling, E.A., 718, 735; 782 Stirrup, H.H., 248; 348 Stober, Q.J., 557, 671, 697, 704; 782 Stoffers, P., 614; 782 See Förstner, U., 617; 756 See Summerhayes, C.P., 586, 643; 782 Stokes, P.M., 548, 575; 782 Stolte, H.A., 197, 199, 212, 219; 348 Stolzberg, R.J., 606; 783 Stoner, J.H. See Chester, R., 635, 638, 643, 644; 751 See Elderfield, H., 754 Stone, J.H. See Spears, L.G., 622; 781

Interlinking of physical

1360

Storch, V. See Welsch, U., 399, 436; 418, 471 Stotzky, G. See Babich, H., 614, 618; 746 Straarup, B.J., 245; 348 See Fenchel, T., 235; 341 Strand, L. See Abrahamsen, G., 202; 337 Straub, C.P. See Williams, L.G., 715; 787 Strause, L. See Case, J.F., 499, 508, 510, 514; 529 Strickland, J.D.H. See Carlucci, A.F., 103; 118 See Fiadeiro, M., 103, 104; 118 See Parsons, T.R., 90; 124 Strohal, P. See Huljev, D., 610; 762 Strom, R.N., 762; 783 Stromgren, T., 547; 783 Strum, J., 503; 533 Strum, J.M., 494, 502, 505, 507; 533 Studentowicz, I., 275; 348 Stuermer, D.H., 609, 610, 615; 783 Stumm, W., 600, 618; 783 Stump, I.G., 667; 783 Sturges, W. See Karl, D.M., 88; 121 Styczynska-Jurewicz, E., 300, 313; 348 Subba Rao, B.V.S.S.R., 219, 227, 228, 253, 277, 280, 297, 313, 322, 323; 341 Subba Rao, N.V. See Ganapati, P.N., 297; 341 Subramanian, T., 514, 615; 783 Suess, E., 84; 126 See Erlenkeu, H., 641; 755 See Müller, P.J., 84; 123 Suffet, I.H., 593; 783 Sugai, S.F., 636; 783 Sugihara, T.T. See Bowen, V.T., 176; 191 Sugimura, Y., 600, 601; 783 See El-Sayed, S.Z., 1; 33 See Miyake, Y., 193 Sugiyama, H. See Naruse, Y., 772 See Matsui, Y., 769 Sugiyama, N. See Haneda, Y., 483, 494, 496; 531 See Johnson, F.H., 531 See Tsuji, F.I., 494; 533 Suitor, J.W. See Ritter, R.B., 622; 777 Sum, F.W. See Andersen, R.J., 612; 745 Sumitra-Vijayaraghavan, M. See Rajendran, A., 540; 776 Summerhayes, C. See Stoffers, P., 614; 782 Summerhayes, C.P., 587, 643; 782 Summers, R.W., 242, 458, 461; 348, 470 Sun, M.S. See Magnuson, V.R., 768 Sunda, W., 551, 579, 583, 600, 603, 724; 783 Sunda, W.B., 603, 604; 783 Sunda, W.G., 550, 551, 554, 559, 603, 621; 783

Oceanography and marine biology See Cross, F.A., 591, 601; 751 See Engel, D.W., 602; 754 Sundararaj, V., 594; 783 Sundby, B. See Bewers, J.M., 783, 617; 747 Surholt, B. See Pörtner, H.-O., 320; 347 Susetai, N. See Hori, K., 467 Sustar, J.F., 628; 783 Sutcliffe, W.H. See Sheldon, R.W., 85, 131, 160; 125, 195 Sutton, D. See Bowen, V.T., 581; 748 Suzuki, H. See Ishii, T., 648; 763 Suzuki, Y. See Sugimura, Y., 600; 783 Sverdrup, H.U., 1, 12, 17, 20; 35 Sverre, O. See Haug, A., 554; 761 Swader, J.A., 551, 603; 783 Swallow, J.C., 40, 63, 72; 73 Swallow, K.C., 612, 617; 783 Swartz, R.C., 628; 783 Swartz, R.W., 109; 126 See Pope, D.H., 110; 124 Swears, S.B. See Phillips, R.R., 453; 469 Sweeney, R.E., 84; 126 Sweetin, R.M. See Amsden, M.P., 592; 745 Swift, D.J.P., 137; 195 Swift, E., 522, 526; 533 Sykes, E.E. See Milanovich, F.P., 658; 770 Symonds, D.J. See Riley, J.D., 422; 470 Syrett, B.C., 590; 784 Syzdek, L.D. See Blanchard, D.C., 108; 117 Szekielda, K.-H., 76; 126 Szeto, C., 551; 784 Tabata,K., 568; 784 See Nishikawa, K., 567; 772 Taborsky, M., 432, 456, 458; 470 Taga, N., 86, 87; 126 Taghorn, G.L. See Feller, R.J., 341 Taguchi, M., 710, 712; 784 Taguchi, S. See Ackley, S.F., 28; 33 Tait, R.I. See Howe, M.R., 41, 48; 73 Takahashi, K., 171; 195 See Baker, E.T., 142; 190 Takahashi, M., 91, 93, 106; 126 Takama, K. See Mizushima, Y., 540; 771 Takei, Y. See Nozaki, I.M., 469 Takita, T. See Kobayashi, T., 439, 448; 467 Talbot, M.S. See Brown, A.C., 358; 416 Talbot,V., 573, 613; 784 Tamblyn, N., 622; 784

887

Interlinking of physical Tamari, Y. See Kusaka, Y., 766 Tamura, E., 426; 470 Tamura, O. See Morii, H., 434; 469 Tamura, S.O., 436; 470 Tanaka, H. See Maeda, M., 602; 783 Tanaka, K. See Ozaki, H., 569; 774 Tanaka, N., 97; 126 Tanaka, Y. See Yamamoto, Y., 788 Taptykova, S.D. See Khovrychev, M.P., 541; 765 Tarifeno, E. See Gordon, M.S., 433; 465 Tarverdieva, M.I. See Permitin, I.U.E., 5; 35 Tasto, R.N., 443, 461; 470 Tatro, M.E. See Girvan, D. C, 592; 758 Tatsumi, Y. See Nozaki, I.M., 469 Tawfik, F. See Harris, J.E., 556; 760 Taylor, A.E.R. See Fox, H.M., 302, 320; 341 Taylor, C.D. See Jannasch, H.W., 113; 121 See Rudd, J.W.M., 106; 124 Taylor, C.W. See King, J.D., 88; 121 See White, D.C., 126 Taylor, D., 588, 594, 642; 784 See Gundersen, K., 119 See Riley, J.P., 602, 637; 777 See Stephenson, R.R., 550, 617; 781 Taylor, D.D., 698, 702; 784 Taylor, D.K. See Magnuson, V.R., 768 Taylor, F. See Windom, H., 787 Taylor, F.E. See Stickney, R.R., 782 Taylor, H.W. See Gordon, A.L., 1, 28; 34 Tchakirian, A. See Bardet, J., 746 Tchernavin, V.V., 525; 533 Tchnernia, P. See Lacombe, H., 57; 73 Teal, J.M., 231, 233; 348 See Backus, R.H., 529 See Burns, K.A., 332; 339 See Smith, K.L., 112, 113; 125 Teare, M. See McLusky, D.S., 264, 300; 345 Tebo, B.M., 484; 533 See Jensen, M.J., 531 See Nealson, K.H., 532 Telek, G. See Eisler, R., 755 Telitchenko, M.M., 540; 784 Templeton, W.L., 570, 580, 600, 659, 742; 783 Tenore, K.R., 241, 333, 334, 336; 348 See Briggs, K.B., 238; 338 See Gardener, W.S., 333; 341 Tessier, A., 600; 784 Tett, P.B., 473; 533

1362

Oceanography and marine biology Thatcher, T.O. See Gibson, C.I., 564; 758 Thayer, G.W., 96; 126 See Wolfe, D.A., 561; 788 Theede, H., 280, 301, 309; 349 Theis,T. L., 590; 784 See Dodge, E.E., 578, 602; 753 Theisen, B. See Bertelsen, E., 488; 529 Thiel, H., 625; 784 See Uhlig,G., 203; 349 Thijssen, R., 459; 470 Thind, K.S., 538; 784 Thines-Sempoux, D. See Zietz-Nicolas, A. M., 507; 534 Thom, R.M. See Armstrong, J.W., 422; 463 Thomas, A.J. See White, G.F., 580, 603; 787 Thomas, D.J., 595, 637; 784 Thomas, J.P., 77; 126 See Hobbie, J.E., 119 Thomas, W.H., 541, 542, 548, 560; 784 Thompson, D.A. See Reynolds, W.R., 643; 777 Thompson, G. See Humphries, S.E., 623; 763 Thompson, J. See Forster, W.O., 637; 756 Thompson, J.A.J., 591, 626, 641, 642, 643; 784 See Olafson, R.W., 550, 567, 582, 584; 774 Thompson, T.G. See Chow, T.J., 636, 637; 751 Thomsen, J.P., 546; 784 Thomson, D.A., 421, 422, 423, 431, 438, 449, 450, 457; 470 See Moffatt, N.M., 437; 468 Thomson, J., 185, 189; 194 See Nozaki, Y., 185; 194 See Turekian, K.K., 185; 197 Thomson, P.A. See Chapman, P.M., 322; 339 Thorell, L., 587; 784 Thorhaug, A., 590; 784 See Ragsdale, H.L., 591; 776 Thorhauge, F., 215; 349 See Jonasson, P.M., 233; 343 Thorndike, E.M., 134, 136; 195 See Eittreim, S., 135; 191 See Ewing, M., 135; 192 See Hunkins, K.E., 135; 192 AA* Thorpe, S.A., 63, 64, 72; 73 Thornton, I., 576, 592, 635; 785 See Aston, S.R., 642; 744 See Halcrow, W., 758 Thornton, L. See Elderfield, H., 755 Thorson, G. See Pearse, J.B., 365; 417 Thrower, S.J., 662; 785 See Ratkowsky, D.A., 776

889

Interlinking of physical

1364

Thun, W. von See Fenchel, T., 263, 289, 334; 341 Thunell, R.C., 142, 171; 195 Thurberg, F.P., 545, 729; 785 See MacInnes, J.R., 406, 544, 620; 416, 768 Thurlow, D.L. See Davis, R.B., 281; 340 Thurston, E.L. See Watson, M., 479, 480; 534 Thurston, J. See Montgomery, J.R., 771 Tiberio, R.D. See McDuffie, B., 770 Tiedge, H. See Meyer-Reil, L.-A., 237; 345 Tiemann, H., 202; 349 Tietjen, J.H., 204; 348, 349 See Alongi, D.M., 236, 281; 337 See Tenore, K.R., 241, 334; 348 Tiffin, L.O., 581, 601; 785 Tikhomirova, A.A. See Morozov, N.P., 785; 771 Timm, T., 203, 204, 207, 269; 349 Timoney, J.F., 548, 643; 785 Timourian, H., 557; 785 Ting, R.Y., 646; 785 Tkachenko, V.N. See Patin, S.A., 540; 775 Toda, S. See Taguchi, M., 712; 784 Todd, E.S., 435, 460; 471 Todd, J.H., 439; 471 Toggweiler, J.R. See Li, Y.-H., 180; 194 Tolksdorf, W., 428; 471 Toole, C.L., 457, 461; 471 Topping, G., 617, 713; 785 See Cambray, R.S., 585; 750 See Saward, D., 712; 778 Tornabene, T.G., 88, 90; 126 Toyodome, T. See Enomoto, N., 661, 665, 670, 672, 681, 686, 695, 699; 755 Tranter, D.J., 1–35 Tranter, D.I. See Smith, D.F., 126 Tranter, H. See Smith, D.F., 126 Trefry, J.H., 591, 626; 785 Treilhard, D.G. See Amsden, M.P., 592; 745 Treni, J.D. See Silver, M.W., 84; 126 Tressler, W.L., 5; 35 Trevallion, A., 355; 418 See Ansell, A.D., 354, 356, 361, 369, 375, 376, 382, 385, 386, 388, Fig. 2 (Brown) facing p. 354, Fig. 7 (Brown) facing p. 376; 415 See Brown, A.C., 351, 388; 415 Trident Engineering Association, 742; 785 Trier, R. See Broecker, W.S., 191 Troughton, J.H. See Christeller, J.T., 84; 118 Trucco, R. See Briand, F., 611; 749 Trueman, E.R., 352, 379, 380, 381, 382, 385, 386, 388, 390, 391, 396; 417 See Jones, H.D., 385; 416

Oceanography and marine biology

891

Tsoukali, E. See Kovatisis, A., 766 Tsubota, H. See Nozaki, Y., 184; 194 Tsuchiyama, F. See Naruse, Y., 772 Tsuji, F.I., 493, 494, 495, 496; 533 See Barnes, A.T., 495, 508, 509; 529 See Haneda, Y., 481, 483, 485, 486, 494, 497; 529 Tsuji, H. See Kusaka, Y., 766 Tsumura, K. See Northcote, T.G., 775 Tsuneki, K. See Nozaki, Y., 469 Tsunogai, S., 181; 195 See Nozaki, Y., 195; 194 Tsuruta, Y., 446; 471 Tsutsumi, T. See Nozaki, I.M., 469 Tsyban, A.V., 108; 127 Tsytsarin, G.V. See Telitchenko, M.M., 540; 784 Tucker, R.E. See Waldhauer, R., 635, 742; 786 Tue, V.T. See Martoja, M., 585; 769 Tugrul, S. See Ramelow, G., 776 Tullis, R.E., 589; 785 Tuncel, G. See Ramelow, G., 776 Turekian, K.K., 185, 539, 585, 591, 593, 595, 600, 642; 195, 785 See Forster, W.O., 637; 756 See Kharkar, D.P., 617; 764 See Nozaki, Y., 185; 194 See Thomson, J., 185, 189; 194 Turner, B.W., 643; 785 Turner, D.R. See Whitfield, M., 601; 787 Turner, J.T., 81, 142, 167; 127, 195 See El-Sayed, S.Z., 1, 27; 33 Turner, L.G.W. See Brown, A.C., 381, 382, Fig. 9 (Brown) facing p. 382; 416 Turner, S.J. See Huppert, J.E., 14, Fig. 8 (Tranter) facing p. 18; 34 Turner, W.D. See Oliff, W.B., 346 Turpaeva, E.P., 549; 785 Turunina, N.V. See Khobot’yev, V.G., 560; 765 Tuttle, J.H., 115; 127 Tusseau D. See Barbier, M., 107; 117 Tyler, A.V., 448, 459; 471 See Wells, B., 458; 471 Tyler, J.E., 134; 195 Tynen, M.J., 236, 239, 261, 275, 283, 297, 305, 307, 313; 349 Tyrrell, D., 88; 127 Uchida, Y. See Enomoto, N., 640; 755 Uchio, T., 624; 785 Ueda, S. See Yamamoto, Y., 788 Uematsu, K. See Ozaki, H., 569; 774 Uemura, M. See Nozaki, I.M., 469 Uglow, R.F. See Price, R.K.J., 546; 776

Interlinking of physical

1366

Uhlig, G., 202; 349 Uhlmann, L. See Zeitschel, B., 154; 196 Ulmer, D.D., 536, 539; 785 Ulitzur, S., 476; 534 Underwood, A.J., 448; 471 U.N. Environmental Programme, FAO., U.N., 550; 784 UNESCO, 593, 617; 785 Ünlü, M.Y. See Small, L.F., 166; 195 Untawale, A.G. See Agadi, V.V., 648; 744 Upadya, G.S., 617; 785 Upitis, V.V., 560; 785 See Pakalne, D.S., 561; 774 Usui, A., 624; 785 Utiger, H., 568; 785 Uysal, H. See Bryan, G.W., 563, 643, 655, 671; 749 See Geldiay, R., 562, 671; 757 Vaccaro, R.F., 91, 558; 127, 785 See Gilliespie, P.A., 553, 558, 605; 757 Vahl, O., 458; 471 See Davenport, J., 428; 464 Valentine, J.W. See Rucker, J.B., 574; 778 van Andel, T.H. See Corliss, J.B., 118 Van Baalen, C. See Armstrong, J.E., 612; 745 Van Boaler, C. See Parker, P.L., 88; 124 Van Boxtel, R. See Yayanos, A.A., 112; 127 Van Buurt, G. See Fonds, M., 427; 465 van de Kamp, G. See Corten, A., Van den Berg, C.M.G., 549, 604; 785 Van den Broek, W.L.F., 242, 549, 683, 694, 700, 701, 704, 712; 349, 785 See Wharfe, J.R., 659, 664, 675, 692, 699, 704, 710; 787 Vanderborght, J.-P., 103, 105; 127 Van der Horst, G. See McLachlan, A., 352, 368, 369; 417 Vanderstappen, R. See De Clerck, R., 712; 753 Van der Weijden, C.H., 592; 785 Van Eck, G.T.M. See Duinker, J.C., 613; 753 Van Grieken, R. See Decleir, W., 537; 752 Van Hoek, R.J. See Spronk, N., 563; 782 Van Hoeyweghen, P. See De Clerck, R., 712; 752 Van Hoven, W., 277, 317; 349 Van Londen, A.M. See De Wolf, P., 622; 752 Van Luik, A.E., 604; 785 van Noort, G.J. See Cruetzberg, F., 446; 464 Van Olst, J.C. See Ford, R.F., 756 Van Vleet, E.S., 88; 127 See Williams, P.M., 107; 127 Varanka, I. See Salanki, J., 543; 778 Varenko. N.I., 570, 714; 786

Oceanography and marine biology

893

Varentsov, I.M., 625; 786 Varini, P.G. See Baudo, R., 592; 746 Vasconcelos, A.C. See Lehman, J., 540, 717; 766 Veerayya, M. See Murty, P.S.N., 624; 772 Vega, V.R. de See Lowman, F.G., 767 Veith, W.J., 441; 471 Vejdovsky, F., 253; 349 Veldhuis, C. See Fonds, M., 430; 466 “Venice System”, Final Resolution, 273; 349 Venkata Ratnam, D. See Subba Rao, B.V. S.S.R., 348 Venkateswara Rao, T. See Subba Rao, B. V.S.S.R., 227, 228, 277, 280, 322, 323; 348 Verigena, I.A., 458; 471 Vermeer, K., 648, 650, 657, 666, 684, 715; 786 See Ohlendorf, H.M., 568; 772 Vernberg, F.J., 539; 786 See Vernberg, W.B., 300, 394; 349, 417 Vernberg, W.B., 300, 311, 395; 349, 418 See DeCoursey, P.J., 591; 752 See Vernberg, F.J., 538; 785 Verriopoulos, G. See Moraitou-Apostolo-poulou, N., 548, 565, 719, 728; 771 Vestergaard, P., 437, 439, 580, 588; 471, 786 Viarengo, A., 544, 574; 786 Vicente, N., 569; 786 Vigers, G.A. See Chapman, P.M., 589; 750 Vikhrenko, N.M., 138; 195 See Klenova, M.V., 138; 197 Villa, Jr, O., 643; 785 Vinogradov, A.P., 642, 650, 654, 660, 663, 666, 670, 671, 673, 675, 677, 680, 686, 689, 693, 694, 696, 697, 699, 701, 703, 704, 706, 708, 710, 711; 786 Vinogradova, N.G., 286; 349 Vinogradova, Z.A., 645; 786 Virnstein, R.W., 228, 242, 243, 280, 283, 290, 323; 349 See Boesch, D.F., 322; 338 Vivian, C.M.G., 588; 786 Vivien, M., 421, 422, 423; 471 Vladimirskaia, E.V., 29, 30; 35 Vlaemink, A. See Decleir, W., 536; 752 Voisin, G., 786 Volkman, J.K., 89; 127 See Perry, G.J., 88; 124 von Damm, K. See Nozaki, Y., 180; 194 Voronina, E.A., 547, 569, 738, 740; 786 Voronina, N.M., 29; 36 Vosjan, J.H. See Boon, J.J., 89; 117 Vuceta,J., 618; 786 Vyncke, W. See De Clerck, R., 712; 753 Wachs, B., 261, 296, 323; 349

Interlinking of physical Wada, E. See Seki, H., 111; 125 Wade, T.L. See Duce, R.A., 754 Wadley, V.A. See Rainer, S.F., 408; 417 Wafar, M.V.M. See Rajendran, A., 540; 776 Wagemann, R., 592, 643; 786 Wagner, F.S., 601; 786 Wagner, G., 238, 261; 349 Waitz, Z., Jr, W.H. See Ackermann, G.R., 587; 744 Wakeman, T.H. See Sustar, J.F., 628; 783 Waksman, S.A., 559; 786 Walden, B. See Smith, K.L., 126 Waldhauer, R., 742; 786 Waldichuck, M., 538, 714; 786 Walker, A.P., 610; 786 Walker, G., 546, 549, 585, 682; 786 Walker, J.G., 280; 349 Wallace, G.T., 108; 127 See Duce, R.A., 753 Wallace, Jr, G.T., 586, 597, 606, 639; 785 See Thomas, W.H., 783 Wallace.J. C., 431; 471 Walsh, J.J., 26; 36 Walsh, P.R. See Duce, R.A., 585; 754 Walter, M.A. See Reeve, M.R., 544, 556; 777 Walters, J.F. See Young, R.E., 525; 534 Walters, J.K. See Jaffe, D., 641, 642, 643; 763 Walther-Mauruschatt, A. See Romesser, J. A., 125 Wallon, A. See Yeats, P.A., 125; 788 Wampler, J.E. See Tsuji, F.I., 534 Wang, C. See Chen, K.Y., 627; 750 Wang, T.S. See Chang, K.-H., 422; 464 Warner, J. See Case, J.F., 523; 529 Warner, J.A., 496, 507, 523, 525; 534 Warnes, J. See McLusky, D.S., 321; 345 Warnes, J.M., 244; 349 Warwick, R.M., 233, 240, 319; 349 Wass, M.L. See Boesch, D.F., 323; 338 Wasserman, A., 583; 786 Watanabe, M. See Matsui, Y., 769 Watchmaker, G. See Timourian, H., 557; 785 Water and Air Res. Inc., Gainsville, Fla, 785 Waterfall, C.E. See Peden, J.D., 578; 775 Watling, H. See Thornton, L, 576; 785 Watling, L., 210, 220, 221, 226, 270; 349 Watson, M., 479, 480, 483; 534 Watson, S.W., 78; 127 Watzin, M.C. See Bell, S.S., 282; 338 Wavre, M., 237, 240, 283; 349

1368

Oceanography and marine biology

895

Wayenbergh, M. See Boeyé, A., 106; 117 Wearn, R.B., 20, 25; 36 Weast, R.C., 143; 195 Webb, J.S. See Elderfleld, H., 755 Webb, K.L., 78; 127 See Crawford, C.C., 90; 118 Webber, H.H., 369; 418 Weber, L.J. See Andersen, P.D., 619; 745 Weber, R.E., 317, 320; 349 Wedborg, M. See Dryssen, D., 349; 754 Weigel, H.P., 554, 635, 639; 786 Weiland,A. L., 537; 786 Weise, G. See Auerbach, S., 554; 745 Weisel, C.P. See Piotrowicz, S.R., 586; 776 Weiss, C.M., 547, 622, 723, 724, 726, 728; 786 Weiss, R.F. See Krishnaswami, S., 193 Weitzman, S.H. See Greenwood, P.H., 474; 530 Welch, P.S., 275, 283, 297; 349 Wellerhaus, S. See Krey, J., 193 Wells, A.W., 425, 442, 457; 471 Wells, B., 459, 461; 471 Wells, J.B.J. See Munro, A.L.S., 334; 346 Welsch, U., 399, 436; 418, 471 Welsh, H. See Peter, R., 471; 775 Wenk, A. See Gieskes, J.M., 40; 72 Wentworth, J.W. See Albright, L.J., 558; 744 Wenz, W., 352; 418 Wenzloff, D.R., 571, 661, 672, 689; 787 See Greig, R.A., 567, 580, 699, 701, 704, 711; 758 Werringloer, J., 539; 787 Wert, M.A. See Stober, Q.J., 557; 782 Westall, J.C., 552, 608; 787 See Swallow, K.C., 611; 783 Westerlund, S. See Magnusson, B., 635 637; 768 Westheide, W., 205, 209; 350 Westheide, W. See Meineke, T., 223, 259 260; 345 Westlake, G.F. See Kleerekoper, H., 567; 765 Westphal, D., 242; 350 See Schmidt-Moser, R., 243; 347 Westrick, J.D. See Theis, T.L., 590; 784 Wharfe, J.R., 264, 323, 326, 333, 661, 666 677, 694, 701, 705, 710; 350, 787 Wheeland, K.G. See Delisle, C.E., 582, 602; 753 Wheeler, A., 422, 445; 471 Wheeler, R.B., 588; 787 Wheeler, R.S. See Marvin, K.. T., 626; 769 Whitton, B.A. See Harding, J.P.C., 551; 760 White, A., 537; 787 White, D. See Windom, H., 787

Interlinking of physical

1370

White, D.B. See Stickney, R.R., 782 White, D.C., 88; 126 See Bobbie, R.J., 90; 116 See Fazio, S.D.,79; 118 See King, J.D., 87; 121 White, G.A. See Smith, K.L., 114; 126 White, G.F., 580, 603; 787 White, M. See Frazer, J.Z., 757 Whitear. M., 432; 471 Whitfleld, M., 601; 787 Whitfield, P.H., 550, 552, 605; 787 See Lewis, A.G., 537, 601, 743; 766 Whitfield, Jr, W.K., 626; 787 Whitley, L.S., 237, 324; 350 Wichman, H.I. See Hiltner, R.S., 636; 761 Widdows, J., 556; 787 Wiebe, P.H., 81, 82; 127 Wiebe, W.J., 86, 87, 97, 155, 166; 127, 195 See Christian, R.R., 110; 117 See Hobbie, J.E., 119 Wiederholm, T., 202, 203; 350 Wiener, G. See Herbert, J.G., 714; 761 Wiener, J.G., 590; 787 Wiese, J., 325; 349 Wieser, W., 260, 274, 311, 714; 349, 787 See Dallinger, R., 714; 754 See Gnaiger, E., 274; 341 Weiss, A. See Silver, S., 550; 780 Wiggins, E.A. See Rainbow, P.S., 574; 776 Wigham, T. See Owens, A., 429; 469 Wijant, J. See Billen, G., 90; 116 Wildung, R.E. See Schmidt, R.L., 579; 779 Wilkes, D.J. See Leland, H.V., 536, 600; 766 Wilkes, F.J. See Reish, D., 777 Wilkie, D.W., 445, 447, 456; 471 Willey, J.D., 616, 643; 787 Williams, A.J. See Staresinic, N., 159; 195 Williams, D. See Corliss, J.B., 117 Williams, G.C., 451; 471 Williams, L.G., 715; 787 Williams, N.V., 203, 217, 317, 323; 349 Williams, P.J.Le B., 76, 90, 91, 92, 94, 95, 97, 101, 116; 126, 127 See Andrews, P., 90, 92; 116 See Banoub, M.W., 90, 92; 116 See Derenbach, J.B., 94, 97; 118 Williams, P.J.L. See Moore, R.M., 118; 771 Williams, P.M., 76, 84, 88, 106, 107, 111, 588, 601, 607; 127, 787 See Carlucci, A.F., 108; 117

Oceanography and marine biology

897

See Rosen, W., 612, 618; 777 Williams, R., 77, 264; 127, 350 See Lindley, J.A., 78; 122 Williams, R.B., 715; 787 Williams, R.J. See Bebbington, G.N., 746 See MacKay, N.J., 566; 768 Williams, R.P.J., 581; 787 Williams, S.C. See Olsen, C.R., 774 Williams, S.T. See Gray, T.R.G., 94; 119 Willis, J.N. See Cross, F.A., 581; 752 See Duke, T.W., 640; 753 Willis, P.-J. See Sieburth, J.McN., 126 Willkon, H. See Erlenkeu, H., 642; 755 Wilson, D.E., 593, 601, 605, 610; 787 See Peden, A.E., 422, 423; 469 Wilson, D.W. See Milanovich, F.P., 550; 770 Wilson, E.M. See Albright, L.J., 558; 744 Wilson, L. See Frazer, J.Z., 757 Wilson, R.C.H., 568; 787 Winchester, J.W. See Pilotte, J.O., 776 Windley, B.F., 575; 787 Windom, H., 693, 694, 696, 697, 699, 702, 703, 706, 708, 710; 787 Windom, H.L., 561, 574, 575, 579, 586, 591, 628, 630, 635, 643, 657, 663; 787 See Stickney, R.R., 782 See Topping, G., 617; 784 Winfrey, M.R., 106; 127 Wing, A.S. See Backus, R.H., 528 Wing, S.S. See Syrett, B.C., 590; 784 Wingert, R.C., 438, 443; 471 Winget, C. See Wiebe, P.H., 81, 166; 127, 195 Winget, C.L. See Jannasch, H.W., 110; 121 Winner, R.W., 557; 788 Wirsen, C.O., 110, 112; 127 See Jannasch, H.W., 110, 111, 112, 114; 121 See Karl, D.M., 114; 121 Wirtz, P., 439, 440, 453; 471 Wisely, B., 552, 562, 724, 725, 734; 788 Wiseman, S.E. See Reynolds, C.S., 154; 194 Wiśniewski, R.J., 214, 219; 350 Witters, R. See Lontie, R., 612; 767 Wittmann, G., 716; 788 Wium-Anderson, S. See Steemann Nielsen, E., 540, 560, 578, 603, 715, 720, 722; 782 Wolf, H.W. See McKee, J.E., 540, 742; 770 Wolf, P.L., 576, 579, 628, 656, 657; 788 Wolfe, D.A., 561, 570, 588; 788 See Duke, T.W., 640; 753 Wolfe, R.S. See Romesser, J.A., 125 Wollast, R. See Vanderborght, J.-P., 103, 105; 127

Interlinking of physical Won, J.H., 642; 788 Wong, C.S. See De Mora, S.J., 623; 753 Wong, K.M. See Bowen, V.T., 191 Wong, M.H., 552, 636, 643, 678, 684; 788 Wong, P.T.S., 742; 788 See Van den Berg, C.M.G., 550; 785 Woo, C.C. See Doyle, L.J., 584; 754 Woo, N.Y.S. See Fryer, J.N., 429; 465 Wood, E.D., 643; 788 Wood, E.J. See Bannister, W.M., 536; 746 Wood, G.F. See Hura, M., 539; 762 Wood, J.M., 539; 788 Wood, L.W., 281; 350 Wood, P.C., 556; 788 Woods, K.R. See Engle, R.L., 536; 755 Wooldridge, T. See McLachlan, A., 352, 417 Woolkalis, M.J. See Baumann, P., 476; 528 Woolmington, A.D. See Davenport, J., 437; 465 Woolner, L. See Riley, J.D., 422; 470 Wootton, R.J. See Ireland, M.P., 578, 675, 679; 763 Word, J.Q. See Reish, D.J., 587; 777 Wort, D.J., 577; 788 Worthington, L.V., 59; 73 Wourms, J.P., 438,385; 471 Wright, B. See Goulder, R., 559; 759 Wright, D.A., 682, 683, 689, 696, 698, 705, 712; 788 Wright, P.G. See Denton, E.J., 511, 520; 530 Wright, R.T., 91, 97; 128 Wright, W.G., 427, 436; 471 Wright, W.R. See Worthington, L.V., 59; 73 Wu, D.C. See Hung, T.C., 588, 659; 763 Wu, M.F. See Yew, D.T., 426; 471 Wu, W.L. See Chang, K.-H., 422, 449; 464 Wuycheck, J.C. See Cummins, K.W., 233, 241; 340 Wyatt, J.T., 742; 788 Wyrtki, K., 9, 14; 36 Xavier, A., 624; 788 Yablonskaya, E.A., 243; 350 Yaffe, P.L., 553; 788 Yager, C.M., 581; 788 Yamagata, S. See Naidu, A.S., 611; 772 Yamaguchi, Y. See Seki, H., 91; 125 Yamamoto, J. See Matsumoto, T., 769 Yamamoto, S. See Zirino, A., 601, 602; 789 Yamamoto, Y., 575, 643; 788 Yang, S.H., 551, 558, 620, 624; 788

1372

Oceanography and marine biology

899

See Ehrlich, H.L., 623; 754 Yaro, I., 89; 128 Yasuda, K. See Taguchi, M., 712; 784 Yayanos, A.A., 112; 128 See Schwarz, J.R., 111; 125 Yeats, P.A., 125; 788 See Bewers, J.M., 601, 617; 747 Yeh, Y. See Milanovich, F.P., 550; 770 Yentsch, C.S. See Laird, J.C., 194 See Williams, P.J.Le B., 76, 90, 95; 127 Yesaki, I. See Koch, F.A., 588; 765 Yetinson, T., 477; 533 See Shilo, M., 476; 532 Yevich, P. See Doyle, L.J., 584; 754 Yevich, P.P. See Rheinberger, R., 584; 777 Yew, D.T., 426; 471 Yingst, J.Y., 236, 279, 282; 350 Yip, A.Y.-W. See Gordon, M.S., 433, 456; 466 Yoakum, R.L. See Davis, W.P., 592; 752 Yone, Y. See Deshimaru, O., 538; 753 See Sakamoto, S., 537; 778 Yonge, C.M., 376, 420; 418, 471 Yosha, S.F. See Cardeilhac, P.T., 545, 719; 750 Yoshiba, S., 484; 534 See Haneda, Y., 487; 530 Yoshida, Y. See Kurata, A., 559, 643, 720; 766 Yoshinari, T. See Subramanian, T., 514; 783 Yoshiyama, R.M., 444, 457, 459; 471 Young, D.J. See Brown, A.C., 352, 404; 416 Young, D.R., 589, 590, 622, 659, 664, 666, 682, 686, 694, 696, 702, 704, 709, 711, 713, 715; 788 See Alexander, G.V., 556, 564, 658, 664; 743, 744 See De Goeij, J.J.M., 752 See Jan, T.-K., 586; 763 See McDermott, D.J., 567; 769 Young, E.G., 576; 789 Young, J.S., 546, 551, 579, 581, 606, 659, 731, 739; 789 Young, M.L. See Moore, R.M., 789; 771 Young, R.E., 522, 523, 525; 534 Young, R.G., 560; 789 Young, R.J., 613; 789 Yuan, L. See De Vries, A.L., 9, 12; 33 Yuzuriha, M. See Tamura, S.O., 436; 470 Zaba, B.N., 573; 789 Zachary, J.L. See Westall, J.C., 552; 787 Zadeh-Douraghi, K. See Eglinton, G., 89; 119 Zafiriou, O.C. See Jannasch, H.W., 82, 154; 121, 193 Zafiropoulos, D., 647; 789

Interlinking of physical

1374

Zahuranec, B.J., 567; 789 Zajicek, O.T. See Pojasek, R.B., 593; 776 Zama, K. See Mizushima, Y., 540; 771 Zamierowski, E.E. See Bergman, S.C., 593; 747 Zander, C.D., 425, 429, 431, 432, 434, 435, 439, 444, 453, 457, 460, 462; 471 Zaneveld, R.J.V., 136; 195 See Bartz, R., 134; 190 See Kitchen, J.C., 142; 192 See Pak, H., 104, 135; 124, 194 Zanicchi, G. See Capelli, R., 750 See Frache, F., 756 See Viarengo, A., 785 Zarra, K. See Kunimoto, M., 89; 122 Zdybiewska, M., 622; 789 Zehnpfennig, R.G. See Merz, R.C., 612; 770 Zeikus, J.G. See Winfrey, M.R., 106; 127 Zeitoun, M.A., 589; 789 Zeitschel, B., 154, 159; 196 Zenk, W., 40, 45,45, 53, 57, 58, 59, 61, 64, 66, 71, 72; 73 Zeuthen, E., 394; 418 Zeutschel, R.P. See Anderson, G.C., 77; 117 Zevenhuizen, L.P.T.M., 550, 559, 603; 789 Zhidkova, L.B. See Romanov, A.S., 637; 778 Zhirova, V.V., 559; 789 Zhorov, V.A. See Babinets, A.E., 746 See Bezborodov, A.A., 601; 747 Ziegenbein, J., 57; 73 Zietz-Nicolas, A.M., 507; 534 See Baguet, F., 507; 529 Zingaro, R.A., 539, 600; 789 Zingde, M.D., 649, 652, 653, 656, 695, 698, 699, 704, 706, 708; 789 Zingmark, R.G., 540; 789 Zirino, A., 601, 602, 608, 637; 789 Zirino, A.H., 597, 613, 636; 789 Zitko, P., 606; 789 ZoBell, C.E., 85, 106, 111, 114, 237; 128, 350 See Morita, R.Y., 113, 114; 123 See Oppenheimer, C.H., 110; 123 Zook, E.G. See Hall, R.A., 644; 760 Zoutendyk, P. See Day, J.H., 406; 416 Zsolnay, A., 615; 789 Zuckerman, B.M. See Coler, R.A., 237; 340 Zuleta, A. See Ebeling, A.W., 433; 465 Zurstrassen, G. See Zietz-Nicolas, A.M., 507; 534

Oceanography and marine biology

901

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SYSTEMATIC INDEX References to complete articles given in heavy type

Abarenicola borealis, 661 ” hemipodus, 661 ” pacifica, 661 Acanthogobius flavimanus, 462 Acanthopagarus australis, 693 Acanthophora orientalis, 651 ” specifera, 650 Acarida, 245 Acartia clausi, 100, 548, 549, 565, 647, 720, 727 ” longiremis, 566 ” simplex, 727 ” tonsa, 549, 552, 556, 565, 646, 727, 738, 743 Acentrogobius meteori, 444 ” ornatus, 443 Acentrophryne, 475 Achaeta, 205 ” eiseni, 316 Acipenser transmoritanus, 693 Acmaea digitalis, 675 ” scabra, 549, 658, 673 Acropoma, 474 ” hanedai, 485 ” japonicum, 485 Acropomatidae, 474, 478, 485 Actinomyxidia, 254 Actinosa metridium, 660 Adelodrilus, 204 Aeolosoma hemprichi, 204 ” litorale, 204, 225 ” maritimum, 204, 208 ” quaternarium, 208 Aeolosomatidae, 198, 200, 205, 209, 225 Agardiella, 651 Agmenellum quadruplicatum, 612 Agropecten irradians, 584 Agrostis gigantea, 548 Aktedrilus, 198, 204, 289, 310 ” monospermathecus, 206, 207, 212, 214, 215, 223, 224, 225, 229, 230, 236, 247, 259, 260, 261,

Oceanography and marine biology

903

263, 269, 274, 275, 276, 283, 286, 287, 288, 289, 294, 299, 300, 301, 302, 303, 305, 309, 310, 326 Alaria esculenta, 652 Albertia naidis, 254 Alcyonium digitatum, 660 Alisma plantago-aquatica, 558 Allocentrotus, 687 Allonais spp., 209 Alloteuthis subulata, 673 Alma emini, 320 Alosa sapidissima, 693 Alteromonas, 542 ” hanedai, 475, 476 Alticus kirkii, 434, 435 ” saliens, 431, 432, 435, 439, 448 Ammodytes lancea, 693 Ammodytes lanceolatus, 693 ” tobianus, 692 Ampelisca milleri, 661 Amphibola crenata, 675 Amphibolis antarctica, 656 Amphichaeta, 205, 209, 235, 287 ” leydigii, 209, 219, 226 ” sannio, 206, 207, 209, 210, 219, 221, 224, 225, 226, 228, 229, 231, 233, 234, 235, 290, 291, 292, 294, 320, 321, 322 Amphidinium carterae, 541 ” carteri, 619, 720 Amphineura, 673 Amphioxus, 291, 337 Anabaena 7120, 604 ” cylindrica, 611 ” flos-aquae, 611 Anabas testudineus, 693 Anadana subcrenata, 661 Anadara granosa, 562, 732 ” senilis, 543, 731 Anarchicas lupus, 693 Anchoa mitchelli, 693 Anchoviella indica, 693 Ancylopsetta quadrocellata, 694 Anemonia sulcata, 462 Anguilla anguilla, 546, 574, 613, 694 ” rostrata, 694 Anguilliformes, 474 Anistremus davidsoni, 694 Ankistrodesmus braunii, 604 Annelida, 660, 724 Anodonta, 661

Interlinking of physical ” cygnea, 543 Anomalocera patersoni, 565, 647, 648 Anomalopidae, 474, 478, 491 Anomalops, 474, 478, 480, 481, 483 ” katoptron, 477 Fig. 1 (Herring) facing p. 420 Anoplarchus purpurescens, 428, 430, 433 Anoplophrya, 247, 248 ” debaisieuxi, 245 ” elongata, 247 ” filum, 245, 247 ” fusiformis, 245, 247 ” lineata, 245 ” nodulata, 247 ” pachydrili, 247 ” paranaidis, 247 ” polydorae, 247 Anoplophryinae, 247, 249 Anser rossii, 576 Anthea cereus, 660 Anthopleura elegantissima, 660 Antimora, 575 ” rostrata, 575 Aplysia punctata, 675 Apodichthys flavidus, 445, 447, 456 Apogon, 492, 493, 494, 498 ” ellioti, 494 ” ellioti (=A.marginatus), 492 Apogonidae, 474, 478, 485 Arbacia, 566, 727, 740 ” lixula, 686 Archamia, 494 Archidoris britannica, 675 ” montereyensis, 673 ” undosa, 673 Archigetes, 254 Archosargus probatocephalus, 545, 719 Arctica islandica, 571, 659, 661 Arenicola, 320, 368 Argentina silus, 694 Argyropelecus, 497, 499, 511, 514, 515, 519, 522, 523, 526 ” aculeatus, 522 ” affinis, 519, 522 ” hemigymnus, 511, 512, 694 Arripis georgianus, 694 ” trutta, 694 Artedius lateralis, 444, 446, 457 Artemia, 82, 544, 564 ” salina, 82, 548, 556, 564, 717, 719, 728, 739

1378

Oceanography and marine biology Arthrobacter, 100, 612 ” marinus, 558, 579, 580 Ascidacea, 738 Ascidia nigra, 738 Ascophyllum, 648 ” nodosum, 283, 561, 575, 576, 580, 651, 713 Astacidae, 565 Asterias rubens, 687 Asteroideae, 687 Astomata, 246, 249 Atelecyclus septemdentatus, 682 Atherinops affinis, 694 Attheya decora, 541, 721 Aulacomya, 362 Aulodrilinae, 205 Aurelia, 660 Axoclinus, 443 Azobacteri, 720 ” vinelandii, 100 Azotobacter, 558 Bacescuella, 204 Bacillariophyceae, 645 Bagre marinus, 694 Bairdiella chrysura, 694 ” icistria, 694 Balanus aburneus (sic), 682 ” amphithrite, 681, 728 ” amphitrite niveus, 544 Balanus balanoides, 549, 585, 682, 715 ” eburneus, 555 ” improvisus, 681, 728 ” perforatus, 681 Barnea dilatata japonica, 661 Bathybembix, 675 Bathydrilus, 204 ” hadalis, 287 ” rarisetis, 290 Bathygobius soporator, 451 Bathylagidae, 474, 478, 488 Batoidea, 367 Beggiatoacea, 116 Belone belone, 695 Beneckea, 476, 483 ” harveyi, 475, 476, 487 ” (=Lucibacterium) harveyi, 476 ” splendida, 475, 476 Bermudrilus, 204

905

Interlinking of physical

1380

Beröe cucumis, 647 Beryciformes, 474, 478 Blenniidae, 442, 443, 454 Blennius, 425 ” cristatus, 441, 454 ” fissicornis, 425 ” pavo, 428, 436, 439 ” pholis, 423, 428, 430, 432, 433, 434, 435, 441, 442, 449, 453, 454, 455, 457, 460 ” rouxi, 438, 439 ” sanguinolentus, 432, 455, 457, 458 ” sphinx, 428, 438, 439 ” tentacularis, 438 Boleophthalmus boddaerti, 436 ” chinensis, 435, 454 ” dussumierei, 438, 441, 457 ” pectinirostris, 434, 694 Borophryne, 475 Bothidae, 455 Brachiomonas submarina, 541, 721 Brachionus plicatilis, 727 Brachydanio rerio, 547, 740 Brachydontes variabilis, 661 Branchiura sowerbyi, 277 Briarum, 660 Brissopsis, 687 Brosmiculus, 474, 488 Bryopsis, 649 Bryozoa, 660, 726 Buccinacea, 351, 365 Buccinanops, 351, 352, 358, 375 Buccinidae, 358 Buccinum, 351, 379, 399 ” undatum, 675 Bugula neritina, 557, 726 Bullia, 351–418; 351, 352, 354, 355, 357, 358, 362, 365, 366, 367, 368, 369, 373, 375, 376, 379, 380, 381, 382, 385, 386, 388, 390, 391, 395, 396, 397, 399, 400, 403, 404, 405, 406, 407, 408, 409, 410, 414, Fig. 18 (Brown) facing p. 408 Bullia anulata, 354 ” callosa, 355 ” digitalis, 352, 355, 357, 358, 362, 365, 366, 367, 368, 369, 370, 371, 372, 373, 374, 375, 379, 381, 382, 385, 386, 388, 389, 390, 391, 392, 393, 394, 396, 397, 399, 400, 402, 403, 404, 405, 407, 408, 409, 410, 414, Fig. viii (Brown) facing p. 352, Fig. 9 (Brown) facing p. 382, Fig. 15 (Brown) facing p. 397, Fig. 18 (Brown) facing p. 408 ” diluta, 354, 368 ” laevissima, 352, 354, 358, 365, 366, 367, 368, 370, 371, 372, 379, 381, 386, 406, 408 ” livida, 375, 376 ” melanoides, 351, 352, 354, 358, 366, 369, 370, 375, 376, 386, 388, 393, Fig. 15 (Brown) facing p. 354, Fig. 7 (Brown) facing p. 376

Oceanography and marine biology

907

” miran, 376 ” natalensis, 354 ” pura, 352, 354, 357, 366, 367, 368, 370, 371, 372, 373, 374, 388, 389, 390, 409 ” rhodostoma, 352, 354, 355, 357, 315. 363, 365, 366, 367, 368, 369, 370, 371, 372, 373, 374, 375, 376, 381, 386, 388, 389, 390, 399, 400, 402, 406, 407, 408, 409, Fig. 4 (Brown) facing p. 358, Fig. 10 (Brown) facing p. 382, Fig. 18 (Brown) facing p. 408 ” tenuis, 354, 375 ” tranquebarica, 375, 376 ” vittata, 354, 369, 374 Busycon canaliculatum, 675 Calanus, 100, 544 ” pacificus, 78, 100 ” plumchrus, 729 Calidris alpina, 244 ” canutus, 244 Caliostoma zyzyphinium, 675 Calianassa, 362, 367, 369, 436 ” californianus, 681 Callinectes sapidus, 538, 556, 682 ” similis, 564, 729 Callorhynchus capensis, 367 Calyptraeidae, 376 Campanularia flexuosa, 548, 726, 739 Cancer anthonyi, 682 ” irroratus, 545, 729 ” magister, 681 ” pagurus, 681 ” productus, 681 Candida utilis, 541, 581, 620 Capitella capitata, 208, 273, 321, 547, 724 Caranx georgianus, 694 Carcharhinus falciformis, 695 ” milberti, 694 Carcharhinus obscurus, 695 Carcharias sorrakowah, 695 Carcinus maenas, 243, 537, 545, 574, 682, 683, 714, 729 Cardium edule, 550, 662, 732 Carex lyngbyei, 656 Caridina rajadhari, 545 Caryophyllaeus, 254 Cassiopea xamachana, 660 Catestomus macrocheilus, 695 Caulerpa peltata, 649 ” racemosa var. laetevirens, 649 ” serularioides, 649 Caulophrynidae, 474 Cebidichthys violaceous, 433, 435, 436, 457

Interlinking of physical

1382

Centronotus (Pholis) gunnellus, 441 Centrophryne, 475 Centrophrynidae, 475 Centropristes striatus, 695 Cephalopoda, 673 Cepphus columba, 461 Ceramium pedicellatum, 561 Cerastoderma edule, 662 Ceratias, 474, 489 Ceratiidae, 474 Ceratoscopelus maderensis, 695 ” townsendi, 509, 510 ” warmingii, 694 Cernosvitoviella, 205 ” immota, 247, 287, Figs. 8 and 9 (Giere & Pfannkuche) facing p. 248 Cetonurus, 474 Chaenophryne, 474, 489 ” draco, 489 Chaetogaster, 276 ” diaphanus, 209, 292 ” diastrophus, 292 Chaetognatha, 647, 727 Chasmichthys, 450 ” dolichognathus, 424, 442, 449, 459 ” gulosus, 424, 442, 449, 459 Chasmodes bosquianus, 437, 440, 447, 453, 461 Chauliodus, 511, 514, 515, 519 ” sloani, 512, 519, 522, 694 Chimaera monostra, 695 Chione stutchburyi, 676 ” undatella, 675 Chironomus, 459 ” tentans, 511, 578 Chlamydomonas, 739 ” reinhardii, 541 Chlamys opercularis, 550, 578, 662 Chlorella, 541, 548, 554, 575, 583 ” pyrenoidosa, 720 ” stigmatophora, 541, 554 ” vulgaris, 611, 720 Chlorophthalmidae, 474, 478, 488 BB Chloróphthalmus, 474, 526 ” albatrossi, 488 ” nigromarginatus, 488 Chlorophyceae, 646, 649 Choanites ficus, 660 Chondrus crispus, 577, 651 Chorda filum, 653

Oceanography and marine biology

909

Chordaria flagelliformis, 653 Chordata, 689 Choromytilus, 362 ” meridionalis, 362 Chromis punctipinnis, 477 Chrysophrys auratus, 696 ” guttulatus, 694 ” major, 537 Chthamalus stellatus, 683 Ciliata, 246, 248 Ciliata, 436 ” mustela, 435, 436 Ciono intestinalis, 615, 646, 689 Cirriformia spirabranchia, 658 Citharichthys sordidus, 696 ” spilopterus, 696 Cladophora, 239 ” fascicularia, 649 Cleidopus, 474, 484, 491 ” gloria-maris, 483 Clevelandia, 461 ” ios, 421, 424, 425, 438, 439, 452, 461 Clinocottus acuticeps, 434, 437 ” analis, 430, 435, 443, 444, 446, 449, 456 ” globiceps, 449, 450 ” recalvus, 435, 436 Clinus cottoides, 457 ” superciliosus, 440, 453 Clitellinae, 12 Clitellio, 205 ” arenarius, 206, 207, 226, 229, 233, 244, 245, 254, 258, 269, 292, 293, 313, 321 Clupea harengus, 547, 567, 569, 696, 735 ” harengus pallasi, 735 ” pilchardus, 696 ” sardinea, 696 ” sprattus, 696 Cnidaria, 359, 362, 726 Coccochloris elabans, 721 Codium elongatum, 649 Codokia orbicularis, 577 Coelenterata, 647, 660 Coelogynopora schulzii, 245 Coelorinchus, 474, 487 Cognettia sphagnetorum, 217 Colpomenia sinuosa, 653 Columbellidae, 375 Cominella adspersa, 676 Conger, 696

Interlinking of physical

1384

Coracinus capensis, 367 Corallina officinalis, 651 Coralliodrilus, 198, 204 Coralliozetus, 443 Corbicula fluminae, 563 Corophium, 459, 461 ” acherusicum, 729 ” volutator, 729 Coryphoblennius galerita, 440, 441, 454, 456, 457, 462 Coscinodiscus curvisetus, 645 ” jonesianus, 645 Cottidae, 440, 443, 454 Cottus asper, 696 ” bubalis, 427 ” scorpius, 427 Crangon allmani, 683 ” crangon, 243, 545, 729 ” vulgaris, 548, 578, 683 Crassostrea angasi, 662 ” angulata, 661 ” commercialis, 661, 688 ” gigas, 543, 545, 546, 575, 576, 610, 661, 688, 719, 731, 739 ” glomerata, 663 ” madrasensis, 562, 663, 688, 731 ” rhizophora, 576 ” virginica, 544, 545, 551, 555, 562, 576, 578, 584, 618, 658, 663, 718, 732 Crepidula fornicata, 676 Cricosphaera elongata, 606, 721 Crocethia alba, 367 Crustacea, 5, 647, 681, 727 Cryptochiton stelleri, 673 Cryptopsaras, 475, 489 ” couesi, 523 Ctenophora, 647, 727 Cyanea capillata, 647 Cyanophyceae, 651 Cyclonassa donovani, 371 ” neritea, 371 Cyclopterus lumpus, 696 Cyclotella nana, 722 Cyclothone, 518 ” acclinidens, 518, 525 ” microdon, 518 ” obscura, 518 ” signata, 525 Cylindrotheca clostridium, 540, 717, 721 ” ” var. californica, 540 Cymatogaster aggregata, 697

Oceanography and marine biology Cynoscion atelodus, 697 ” nebulosus, 696 ” regalis, 696 ” xanthulus, 696 Cypraea spadacea, 676 Cypridina, 485, 492, 493, 494, 495, 496, 497, 498 ” hilgendorfii, 492, 493, 495, 496 ” noctiluca, 494 Dalyelloidea, 245 Daphnia, 564 ” magna, 564, 601, 717 ” magnum, 729 Dasya, 651 Delesseria sanguina, 651 Dero, 209 ” limosa, 233 ” nivea, 276 Deschampsis cespitoma, 656 Desulphovibrio, 89 Diademichthys lineatus, 462 Dialommus fuscus, 435, 460 Diaphus, 496, 497 ” holti, 510 ” rafinesquii, 522 Dibranchus, 496 Diceratias, 474 Dictyopteris australis, 653 Dictyosiphon chordaria, 653 ” foeniculaceus, 653 Dictyota, 653 ” bartayresii, 653 ” dumosa, 653 Diogenes brevirostris, 368 Diplodus, 459 ” cervinus, 459 ” sargus, 459 ” vulgaris, 696 Diretmus, 526 Distichlis spicata, 657 Distichus spicata, 657 Ditylum brightwelli, 541, 582 Donax, 356, 367, 369, 371 ” gouldii, 675 ” serra, 367, 368 ” sordidus, 367 Dorsanum, 351, 352, 358, 399, 400 ” moniliferum, 375

911

Interlinking of physical

1386

Dorosoma petenense, 697 Dunaliella marina, 541 ” primolecta, 721 ” salina, 204, 540, 611, 721 ” tertiolecta, 541, 612, 721 Echinarachnius, 687 Echinodermata, 687, 727 Echinometra mathaei, 727 Echinus esculentus, 687 Echiostoma, 499, 511, 526 Ecklonia, 358 Ectocarpus, 542, 555, 653 ” siliculosus, 721 Edriolychnus, 475, 489 Edukemius, 205 ” benedii, 204 Egregia menziesii, 653 Eirene viridula, 555, 726 Eisenia, 661 ” bicyclis, 659 Elasmobranchii, 697 Elminius modestus, 566 Elodea nuttallii, 576 Emerita, 356, 684 ” analoga, 683 Enchytraeidae, 198, 200, 204, 207, 225, 228, 230, 267, 278 Enchytraeus, 203, 205, 247, 253, 287, 319 ” albidus, 197, 201, 206, 207, 217, 218, 219, 224, 225, 229, 236, 238, 239, 243, 244, 247, 252, 254, 259, 260, 263, 267, 268, 272, 274, 275, 282, 283, 287, 296, 297, 300, 301, 302, 305, 307, 308, 310, 312, 313, 316, 318, 319, 320, 322, 325, 326, 543 ” barkudensis, 297, 305, 323 ” bigeminus, 215 ” coronatus, 201, 316, 322 Engraulis encrasicholus, 697 ” mordax, 547, 735 Enophrys bison, 438, 461 Ensis ensis, 663 Enteromorpha, 239, 555, 648, 649 ” flexuosa, 649 ” intestinalis, 649 Entomacrodus vermiculatus, 440 Epinephelus aeneus, 698 ” ergastularius, 697 ” guaza, 697 ” guttatus, 697 ” striatus, 697 Erbopdella octoculata, 244

Oceanography and marine biology

913

Erigone longipalpis, 244 Erpobdell octoculata, 575 Escherichia coli, 101, 558 Etmopterus spinax, 698 Etropus crossotus, 698 ” suratensis, 697 Euchaeta japonica, 556 Eudistylia vancouveri, 579, 659, 718 Euglena, 238 Eulamia melanoptera, 698 Eunecia crassa, 660 ” rissoa, 659 Eupagurus, 684 ” bernhardus, 683 Euphausia, 494, 496, 544 ” krohnii, 646 ” pacifica, 494, 730 ” similis, 494 ” superba, 5, 26, 29, 30, Fig. 4 (Tranter) facing p. 2 Euphausiacea, 5 Eurydice longicornis, 367 Euthynnus alletteratus, 698 Evynnis japonica, 698 Exosphaeroma truncatitelson, 367 Exuviaella, 722 Fridericia, 205, 287, 319 ” bisetosa, 316 ” bulbosa, 257 ” callosa, 287 ” galba, 316 Fucus, 203, 225, 231, 234, 238, 239, 283, 330, 583, 655 ” ceranoides, 653 ” distichus, 437, 648, 653 ” distichus sp. edentatus, 653 ” serratus, 546, 653, 654 ” spiralis, 546, 654 ” vesiculosus, 239, 282, 546, 570, 576, 580, 582, 609, 619, 648, 654, 743 Fugu niphobles, 438 Fundulus, 568 ” heteroclitus, 544, 545, 567, 568, 602, 714, 735 Gadella, 474, 488 Gadiformes, 474, 487 Gadus aeglefinus, 698 ” morrhua, 697 Galathea squamifera, 684 Galeolaria caespitosa, 724

Interlinking of physical Galeorhinus australis, 699 Galeus melastromus, 699 Gamasida, 245 Gammarus, 647 ” pseudolimnaeus, 730 Gastropoda, 675, 735 Gastrosteus aculeatus, 567 Gaussia, 496 Gazza, 474 Gennadas, 496 Genyonemus lineatus, 699 Gibbsonia, 431 ” elegans, 431, 435 ” metzi, 431 ” montereyensis, 431 Gigantactinidae, 475 Gigantactics, 474, 488 Gigartina mamillosa, 651 ” stellata, 650 Gillichthys, 429 ” mirabilis, 429, 461 Girella nigricans, 430, 443, 446, 453 ” tricuspidata, 697 Glenodinium, 722 ” foliaceum, 721 Gloeocystis gigas, 611 Glycymeris glycymeris, 664 Glyptocephalus cynoglossus, 699 Gobiesocidae, 440, 442, 455 Gobiesox meandricus, 457 ” pinniger, 429, 443, 454 ” rhessodon, 424, 441, 456 ” strumosus, 436, 439 Gobiidae, 242, 425, 440, 442, 454 Gobionellus sagittula, 435, 460 Gobiosoma bosci, 437 ” chiquita, 452, 453 Gobius cobitis, 442, 456, 457 ” jozo, 441 ” niger, 455 ” paganellus, 441, 442 Gonichthys, 522 Gonospora pachydrili, 253 Gonostoma, 511, 526 ” atlanticum, 518, 522 ” bathyphilum, 518 ” elongatum, 518 Gonyaulax, 496

1388

Oceanography and marine biology ” tamarensis, 542, 543, 550, 559, 602, 716, 721 Gorgonia acerosa, 660 Gorgonium flabellum, 660 Gracilaria, 651 ” sp. 1, 650 ” sp. 2, 650 ” sp. 3, 650 ” corticata, 650 ” verrucosa, 559, 650 Grania, 198, 205, 228, 257, 261, 263, 287, 291 ” macrochaeta, 258, 290 ” macrochaeta pusilla, 286, 290 ” maricola, 290 ” monospermatheca, 198 ” postclitellochaeta, 226, 228, 275, 290 ” variochaeta, 290 “Gregarina” enchytraei 252 ” saenuridis, 252 Gregarinida, 246 Gryphea angulata, 664 Gymnodinium, 543 ” breve, 625, 626 ” brevis, 721 ” splendens, 721 Gymnogongrus, 651 Halacarellus, 245 Halacarida, 245 Halicryptus spinulosus, 243 Halimeda tuna, 650 Haliotis, 676 ” cracherodii, 552, 675, 734 ” rufescens, 552, 570, 675, 734 ” striata, 675 ” tuberculata, 675 Halmablennius lineatus, 442, 457 Halodule, 657 Halophila, 657 Halosaccion ramentaceum, 651 Harmothoe sarsi, 243 Harpacticoidea, 230 Helix aspersa, 676 Hemifuscus ternatanus, 699 Hemigrapsus, 684 Henrica sanguinolenta, 687 Hesperodrilus litoralis, 205 Heterodrilus, 198, 205 ” arenicolus, 204

915

Interlinking of physical Heteropoda, 647 Heterozostera tasmanica, 657 Hexactinomyxon psammoryctis, 254 Hexagrammos otakii, 699 Hexapanopeus augustifrons, 684 Himantolophidae, 475 Himantolophus, 475, 489 Hinnites giganteus, 664 ” multirugosus, 663, 664 Hipponyx, 376 Histioteuthis, 526 Holocentrus rufus, 699 Holothuria, 688 ” mexicana, 576 Holotricha, 246 Homarus gammarus, 730 ” vulgaris, 683 Hydra littoralis, 555, 726 Hydrobia, 239 Hydroides dianthus, 266 ” parvus, 724 Hydrophilidae, 244 Hygophum benoiti, 699 ” hygomi, 699 Hymenocephalus, 474 Hypleurochilus geminatus, 453, 457 Hypnea, 652 ” muciformis, 651 Hypsoblennius, 439, 440, 446 ” gentilis, 430 ” gilberti, 445, 447, 450 ” jenkinsi, 447 Ichthyococcus, 511, 518 Idiacanthus, 511 Ilyanassa obsoleta, 243 Ilyodrilus templetoni, 277, 297, 299, 317, 323 Ilypnus gilberti, 421, 425, 452 Inanidrilus, 198, 204 Ischadium demissum, 665 Ischnochiton conspicuus, 673 Isochrysis galbana, 541 Istiblennius edentulus, 435, 452 ” zebra, 432, 438, 439, 452 Jamiesoniella, 198, 204 Jania, 652 Jolydrilus, 198

1390

Oceanography and marine biology Joruna tormentosa, 677 Kalyptorhynchia, 245 Kareius bicoloratus, 446 Klebsiella (Aerobacter) aerogenes, 101 ” aerogenes, 100, 101, 559, 602 Krytophanaron, 474, 478, 480, 481 ” alfredi, 477, 480, 481 ” harveyi, 477, 480 Labidocerca scotti, 730 Labyrinthomyxa marina, 553 Lactarius lactarius, 699 Lagodon rhomboides, 553, 590 Laminaria, 648 ” digitata, 609, 654 ” gloustoni, 654 ” hyperborea, 541, 546, 654 ” saccharina, 654 ” stilker, 654 Lamna nasus, 699 Lampanyctus pusillus, 699 Larvacea, 648 Lasiognathus, 475 Lateolabrax japonicus, 699 Lates calcarifera, 700 Laurencia sp. 1, 652 ” sp. 2, 651 Leander adspersus, 684 ” squilla, 683 Leiognathidae, 474, 478, 486 Leiognathus, 474, 486 ” aureus, 486 ” elongatus, 485, 486 ” hataii, 486 ” rivulatus, 485 ” splendens, 699 Leiostomus xanthurus, 700 Lemna perpusilla, 561 Lepidogobius lepidus, 442, 452, 453, 457 Lepidophanes indicus, 700 Lepidorhynchus, 474 ” denticulatus, 487 Leptocottus, 462 ” armatus, 429, 699 Leptocylindricus danicus, 645 Leuresthes sordina, 438 ” tenuis, 437

917

Interlinking of physical

1392

Limacina retroversa, 647 Limanda limanda, 700 ” ferruginea, 699 Limnodrilus, 281, 312 ” claparedeanus, 293, 299 ” hoffmeisteri, 212, 229, 230, 231, 233, 238, 244, 277, 279, 293, 296, 298, 300, 301, 302, 305, 308, 316, 320, 321, 322, 323, 324 ” medioporus, 263, 264 ” profundicola, 293 Limnodriloides, 205 ” barnardi, 265, 290 ” monothecus, 206, 207, 265, 333 ” verrucosus, 265, 333 Limulus polyphemus, 538 Lineus ruber, 244 Lingula unguis, 665 Linophrynidae, 475 Linophryne, 475, 489 ” arborifera, 488 Linophrys adriaticus, 425 ” (Blennius) adriaticus, 443 ” pholis, 436 ” canevai, 424, 443 ” dalmatinus, 424, 443 Lithognathus lithognathus, 367 Littorina littoralis, 677 ” littorea, 549, 574, 578, 584, 612 ” mariae, 460 ” neritoides, 675 ” obtusa, 677 ” rudis, 677 ” saxatilis, 677 Liza macrolepis, 544, 648 Lobianchia dofleini, 700 Loligo opalescens, 700, 673 ” vulgaris, 671 Lophalticus (Alticus) kirkii, 435 ” kirkii, 431, 443, 452 Lophiiformes, 474, 489 Lopholatilus chamaeleonticeps, 700, 712 Lotella, 488 ” phycis, 487 Lottia gigantea, 677 Lucibacterium, 476 ” harveyi, 475 Luidia, 688 Lumbricillus, 198, 205, 236, 238, 240, 242, 243, 244, 247, 252, 254, 267, 287, 290, 318, 336 ” aegialites, 267

Oceanography and marine biology

919

” arenarius, 258, 287 ” bülowi, 257, 258, 287 ” codensis, 263 ” helgolandicus, 287 ” knoellneri, 257, 287 ” lineatus, 197, 201, 203, 206, 207, 217, 218, 219, 224, 225, 228, 229, 230, 231, 234, 236, 238, 239, 242, 245, 247, 252, 254, 258, 259, 260, 263, 267, 268, 274, 275, 279, 280, 282, 283, 284, 287, 296, 297, 305, 306, 307, 308, 310, 312, 313, 316, 318, 319, 322, 323, 326, 327, 328, 329, 333, Fig. 10 (Giere & Pfannkuche) facing p. 252, Fig. 37 (Giere & Pfannkuche) facing p. 329 “lineatus” (?rivalis), 217, 247, 252 ” pagenstecheri, 206, 207, 287 ” reynoldsoni, 236, 260, 283, 296, 297, 305, 307, 312 ” rivalis, 201, 203, 217, 218, 229, 238, 239, 245, 247, 252, 272 287, 296, 297, 301, 312, 316, 317, 319, 322 Lumbricillus semifuscus, 252, 258 ” viridis, 229, 252, 258, 260, 287, 316 Lumbricomorpha, 204 Lunatia heros, 382 Lutjanus blackfordi, 700 ” campechanus, 699 Lymnaea stagnalis, 563 Lyngbya nigra, 543 Lytechinus variegatus, 577, 587 Macoma balthica, 544, 562, 587, 603, 616, 627, 659, 665, 718 ” inquinata, 588 ” modialus, 664 Macrocystis, 553, 560, 655 ” integrifolia, 576 ” pyrifera, 539, 546, 561, 654, 715 Macrouridae, 474, 478, 487 Macruroides, 488 Macrurus rupestris, 700 Macullochella macquariensis, 700 Maia squinado, 684 Makaira indica, 700 Malacocephalus, 474 Malacosteus, 526 Malanogrammus aeglefinus, 700 Mancalais, 488 Margaritifera margaritifera, 571 Marinogammarus marinus, 566, 718, 730 Marionina, 205, 219, 224, 236, 248, 258, 259, 261, 275, 276, 284, 287, 288, 290, 300, 314, 316, 319 ” achaeta, 223, 229, 279, 290, 299, 300, 312, 313, 314, 316, 318 ” crassa, 247 ” elongata, 290 ” mesopsamma, 290

Interlinking of physical

1394

” preclitellochaeta, 215, 287, 290, 299, 301, 305 ” southerni, 206, 207, 225, 236, 247, 258, 260, 267, 283, 287, 299, 300, 305, 308 ” spicula, 206, 215, 223, 225, 229, 230, 258, 259, 260, 268, 287, 290, 299, 300, 313, 314, 316, 318, 323 ” sublitoralis, 286 ” subterranea, 206, 207, 215, 224, 225, 229, 230, 235, 258, 259, 260, 287, 288, 290, 294, 299, 300, 305, 309, 322, 323, 326, 332 ” weilli, 290 ” welchi, 247, 263 Maurolicus, 511, 518 ” muelleri, 699 Meganyctiphanes norvegica, 647, 684 Megascolecidae, 198, 200, 204, 219, 227 Melanocetidae, 474 Melanocetus, 474, 489 ” murrayi, 488 Menidia menidia, 546, 569 Mercenaria mercenaria, 563, 584, 665, 689, 733 Meretrix casta, 562, 733 ” chionae, 664 ” lamarkii, 664 Merlangius merlangus, 242, 701 Merlucciidae, 474, 478, 487 Merluccius bilinearis, 701 ” merluccius, 701 ” vulgar is, 701 Mesenchytraeus gelidus, 268, 275 ” solifugus, 267, 275 Mesniliella, 247, 248, Figs. 8 and 9 (Giere & Pfannkuche) facing p. 248 ” elongata, 247 ” fastigata, 247 Mesobius, 474, 487, 488 ” berryi, 491 Metridia pacifica, 730 Microcystis aeruginosa, 722 Micropogon undulatus, 701, 714 Microscolex, 204 Microstomus pacificus, 545, 567, 701 Mnemiopsis mccradyi, 727 Mnierpes, 433, 456, 462 ” macrocephalus, 433, 435 Modiolus demissus, 544, 563, 733 ” modulus, 543, 563, 664, 732 Mola, 701 Mollusca, 376, 381, 400, 409, 411, 647 Monacanthus chinensis, 701 Monocentridae, 474, 478, 484, 485, 491 Monocentris, 474, 484, 491

Oceanography and marine biology

921

” japonicus, 484, 486, 490 Monochrysis lutheri, 541, 605 Monocystidae, 249 Monocystis, 249 ” lumbricilli, 252 ” pontodrili, 252 ” rhabdoda, 252, Fig. 10 (Giere & Pfannkuche) facing p. 252 Monodonta turbinata, 678 Monopylephorus, 205, 324 ” indicus, 226, 228, 277, 279, 321 ” irroratus, 206, 207, 258, 293, 296, 301 ” parvus, 206, 207 ” rubroniveus, 206, 207, 292, 293, 297 ” waltairensis, 226, 228, 277, 279, 321 Moridae, 474, 478, 488 Morone saxatilis, 702 Mugil cephalus, 590, 702 ” parsia, 701 Mulus barbatus, 702 ” surmulletus, 701 Murex tronchulus, 678 Mustelus antarcticus, 702 ” canis, 701 Mya arenaria, 562, 665, 718, 733, 739 Mycobacterium fortuitum, 101 Mycteroperca phenox, 702 ” tigris, 701 ” venosa, 701 Myctophum, 511, 522 ” obtusirostrum, 523 ” punctatum, 701 ” spinosum, 523, 525 Mylocheilus caurinus, 702 Myoxocephalus scorpius, 702 Mytilaster lineatis, 666 Mytilus, 267, 320, 669, 739 ” californianus, 556, 563, 658, 664 ” edulis, 543, 544, 548, 555, 556, 563, 572, 576, 590, 612, 658, 664, 666, 668, 714, 718, 733, 743 ” edulis planulatus, 556, 563, 668 ” galloprovincialis, 543, 564, 572, 574, 668 ” obscurus Myxocephalus scorpius, 712 Naididae, 198, 200, 205, 206, 208, 227, 229, 269 Naidomorpha, 204 Nais, 219 ” communis, 210, 238, 243, 292 ” elinguis, 204, 206, 208, 209, 219, 226, 229, 233, 235, 237, 242, 254, 269, 276, 287, 292, 322,

Interlinking of physical

1396

326 ” pseudoobtusa, 209, 292 ” variabilis, 210, 254, 292 Nassa pygmaea, 371 Nassariidae, 351–418, 352, 358, 365, 371 374, 375, 399, 408, 411 Nassarius, 351, 352, 357, 358, 367, 374, 392, 407 ” kraussianus, 358, 367, 376 ” obsoletus, 358, 365, 375, 406, 544, 619, 620 ” plicatus, 376 ” reticulatus, 370, 371, 375, 388, 391 ” tiarula, 375 Nassidae, 351 Natica, 367 ” tecta, 367 Navanax merimis, 678 Navicula salinarum, 235 Neanthes, 591, 669 ” arenaceodentata, 581, 724 ” japonica, 543, 724 ” succonea, 668 Nematoda, 230 Nemertini, 230, 245 Neoceratiidae, 474 Neoclinus bryope, 437 Neogardhiella baileyi, 561 Neogastropoda, 411 Neomysis americana, 647 Neoplatycephalus macrodon, 703 Neoscopelus, 497 ” microchir, 496 Nephthys, 661 ” hombergi, 659, 724 Neptunus pelagicus, 684 Nereis, 313, 556, 562, 661, 725 ” diversicolor, 224, 243, 320, 325, 545, 547, 562, 578, 658, 659, 713, 725 ” virens, 243, 659 Nereocystis luetkeana, 577 Nerita tessplata, 577 Neurospora crassa, 574 Nezumia, 474, 487, 488 Nitocra spinipes, 565, 730 Nitzschia closterium, 560, 717, 721 ” palea, 721 Norrissia norrissii, 678 Notoscopelus caudispinous, 703 Nucella lapillus, 678 Odontomacrurus, 474

Oceanography and marine biology

923

Octopus vulgaris, 673 Oikopleura dioica, 100 Oligochaeta, 197–350; 197, 198, 204, 207, 231, 245, 321 Oligochaetocystis pachydrili, 253 Oligocottus maculosus, 424, 429, 443, 444, 445, 447, 449, 450, 451, 455, 460, 461 ” snyderi, 429, 443, 444, 445, 446, 447, 456, 459 Oliva gibbosa, 386 Omnastrephes bartrami, 674 Omobranchus loxozonus, 437 Oncorhynchus kisutch, 545, 736 ” nerka, 537, 701, 703, 712 ” tschawytscha, 703 ” tshawytscha, 568, 735 Oneirodes, 474, 489 ” acanthias, 488 Oneirodidae, 474 “Opalina” (=Anoplophrya) lineata, 246 ” ” pachydrili, 247 “Opalina” filum, 246 Ophichthus gomesi, 703 ” ocellatus, 703 Ophidonais, 209 Ophioblennius atlanticus, 432, 453, 456, 458 Ophiopsila, 496 Ophryotrocha, 273, 564 ” labronica, 564, 725 Opisthopora, 204 Opisthoproctus, 474, 488, 489 Oplophorus, 497 Opsanus tau, 703 Orconectes, 565 ” rusticus, 564 Orthopristus chrysopterus, 704 Ostrea, 670 ” edulis, 548, 610, 644, 668, 670, 741 ” lurida, 670 ” lutaria, 670 ” sinuata, 670 ” virginica, 741 Otolithus ruber, 704 Otoplanida, 245 Ovalipes, 367 ” punctatus, 367 Pachydrilus “lineatus”, 320 Pachygrapsus crassipes, 565, 685 Padina commersonii, 656 ” tenuis, 656

Interlinking of physical

1398

” tetrastromatica, 656 Palaemon elegans, 685, 689 ” serratus, 684 ” squilla, 684 Palaemonetes, 591, 627, 685 ” kadiakensis, 684 ” pugio, 243, 684 ” varians, 684 Palmaria palmata, 652 Pampus argenteus, 704 Pandalus danae, 551, 564, 606, 718, 731 ” montagui, 684 Panulirus interruptus, 549, 685 Paphia, 678 ” luzonica, 677 ” staminea var. laciniata, 677 Paracalanus parvus, 731 Paracentrotus lividus, 557, 566, 688 Paracereis sculpta, 731 Paraclinus integrippinnis, 436 Paragrapsus quadridentatus, 731 Paralabrax clathratus, 704 ” maculatofasciatus, 703 Paralichthys, 590 ” californiens, 703 ” lethostigma, 703 ” olivaceus, 703 Paranais, 205, 219, 242, 276, 287 ” “elongata”, 247 ” frici, 206, 207, 219, 296, 298 ” litoralis, 206, 207, 209, 210, 219, 221, 225, 226, 228, 229, 234, 235, 237, 241, 242, 243, 244, 247, 248, 258, 269, 280, 290, 291, 292, 294, 296, 298, 300, 320, 321, 322 Parapriacanthus, 492, 493, 494 ” beryciformis, 492 ” ransonneti, 492 Paratrachichthys, 474, 483 Paroneirodes, 474 Parophrys vetulus, 457, 704 Parvilux ingens, 525 Patella, 320, 385, 680 ” athletica, 677 ” coerulea, 677 ” intermedia, 612, 677 ” vulgaris (sic), 677 ” vulgata, 574, 612, 670, 677, 679 Patiria, 688 ” miniata, 686 Pecten circularis laquisulcatus, 670

Oceanography and marine biology

925

” fumatus, 670 ” jacobaeus, 670 ” maximum, 578 ” maximus, 670 ” novae-zeolandiae, 670 ” varius, 670 Pectunculus glycemaris, 671 Pelates sexlineatus, 704 Pelecypoda, 732 Pellona ditehela, 704 Peloscolex, 205, 238, 242 ” benedeni, 201, 204, 206, 207, 212, 214, 221, 222, 226, 228, 229, 230, 231, 233, 234, 242, 243, 244, 258, 260, 263, 264, 267, 276, 277, 279, 280, 287, 289, 290, 292, 293, 296, 298, 301, 302, 305, 320, 321, 323, 325, 332 ” gabriellae, 266 ” intermedius, 264, 266 ” multisetosus, 231 Peltorhamphus latus, 455 Pelvetia canaliculata, 547, 656 Panaeus aztecus, 685, 731 ” brasiliensis, 684 ” californiensis, 731 ” douorarum, 731 ” indicus, 684 ” japonica, 684 ” japonicus, 537 ” semisulcatus, 684 ” setiferus, 684 Peosidrilus, 204 Perciformes, 474, 484 Periophthalmidae, 421, 455 Periophthalmus, 441 ” cantonensis, 424, 433, 434, 435, 436, 438, 440, 454 ” chrysospilos, 425, 439 ” dipes, 427 ” expeditionium, 434 ” gracilis, 434 ” kalolo, 439 ” kalolo (=P.koelreuteri), 438 ” koelreuteri, 453, 458 ” sobrinus, 433, 434, 441, 452, 453 ” vulgaris, 435 Perna canaliculus, 671 Phaeocystis, 580 Phaeodactylum tricornutum, 541, 554, 620, 722 Phaeophyceae, 652 Phallodrilinae, 198, 204, 225, 291 Phallodrilus, 204, 263, 287

Interlinking of physical

1400

” aquaedulcis, 294 ” hallae, 204 ” leukodermatus, 226, 228 ” postspermathecus, 198 ” profundus, 287 ” prostatus, 224, 225, 247, 286, 287, 289, 326 Phascalosoma agassizii, 588 Pherallodiscus funebris, 429, 443 Phialidium, 727 Pholidae, 440 pholis, 436 ” gunnellus, 428, 431, 435 ” 435 428, 430 Photobacterium, 475, 476, 483, 489 ” fischeri, 475, 476, 484, 486, 487, 491 ” leiognathi, 475, 476, 483, 484, 486 ” logei, 475, 476 ” mandapamensis, 475 ” phosphoreum, 475, 476, 483, 486, 487, 488, 489, 491 ” (=Vibrio) fischeri, 476 Photobelpharon, 474, 477, 480, 481, 483, 491 ” palpebratus, 477, Fig. 1 (Herring) facing p. 477 Photocorynus, 475 Phreodrilidae, 198, 200 Phyllodoce maculata, 562 Phyllospadix, 445 Physalia, 362, 406, 647, 659 Physiculus, 474, 488 ” rastrelliger, 487 Pilumnus hirtellus, 686 Pisaster, 688 ” giganteus, 686 ” ochraceus, 688 Pitar morrhuana, 577, 659, 671 Placopecten magellanicus, 671 Platicthys flesus, 455, 459, 704 ” stellatus, 544 Platycephalus fuscus, 705 Plectroplites ambiguus, 705 Plesiopora, 204 Pleurobrachia, 544 ” pileus, 727 Pleuronectes limanda, 705 ” platessa, 424, 445, 448, 454, 458, 460, 546, 568, 704, 712, 737 Pleuronectidae, 455 Plexaura flexulosa, 660 ” humomolla, 659 Plumatella fungosa, 660

Oceanography and marine biology

927

Plutellus, 204 Poccillopora, 660 Podocerus fulanus, 732 Poecilia latipinna, 705 Polinices, 352, 381, 385 Polinices duplicatus, 382, 394 Pollicipes polymerus, 686 Polychaeta, 243, 724 Polyipnus, 497 Polypriori americanus, 705 Polypus bimaculatus, 680 Polysiphonia, 723 Pomatomus pedica, 705 ” saltatrix, 575, 704, 711 Pomatoschistus, 242, 437, 461 ” microps, 242, 243, 425, 427, 430, 431, 438, 441, 459, 460, 545 ” minutus, 424, 425, 427, 428, 430, 441, 454, 455, 458, 459, 460, 704 ” pictus, 427, 455 Pontella meadii, 81 Pontodrilus, 204 ” bermudensis, 219, 226, 252, 297, 312, 313, 323 Pontoporeia, 243 Porania pulvillus, 688 Porcellana platycheles, 686 Porichthys, 493, 494, 495, 496, 497, 498, 499, 502, 503, 507, 511, 514 ” myriaster, 503 ” notatus, 493, 495, 503, 508 ” porosissimus, 493 Porifera, 659 Porphyra, 538, 652 ” purpurea, 651 ” umbilicalis, 651 Portunus depurator, 686 ” holsatus, 684 ” puber, 684 Potamalosa novae-hollandiae, 706 Potamothrix hammoniensis, 233, 293, 299, 322, 324, 326 ” moldaviensis, 299 Prionace glauca, 706 Pristina, 209 ” longiseta, 208 Proiana pulvillus Prorocentrum micans, 723 ” minimum, 611 Proseriata, 245 Prosobranchiata, 411 Prosopora, 204 Protothaca staminea, 671, 680

Interlinking of physical Protozoa, 720 Proxenetes, 245 Psammoryctides barbatus, 299, 322 Pseudocalanus, 544, 566 Pseudodiaptomus coronatus, 647 Pseudograffila arenicola, 245 Pseudomonacanthus ayraudi, 706 Pseudomonas cuprodurans, 558, 580 Pseudopleuronectes americanus, 544, 575, 706 Pseudoplexaura crassa, 660 Pterocladia, 444 ” pinnata, 651 Pteropoda, 647 Ptychocheilus oregonensis, 706 Pugettia producta, 716 Purpura lapillus, 574, 613 Pyrosoma, 490 Pyura, 362 Quietula y-cauda, 425, 452 Radiophrya, 248 ” grandis, 245 ” pachydrill, 247 ” prolifera, 247 Raja eglanteria, 706 Rangia, 671 ” cuneate (sic), 591 Rangia cuneata, 556, 627, 671, 734 Rastrelliger, 706 ” kanagurta, 706 Regificola grandis, 707 Renilla, 496 Reporhamphus australis, 707 Rhabdamia, 494 Rhabdosargus holubi, 367 Rhinobatis, 367 ” lentiginous, 706 Rhinoptera bonusus, 707 Rhizophora mangle, 577 Rhizoselenia calaravis, 645 Rhizostoma pulmo, 660 Rhodotorula rubra, 739 Rhodymenia palmata, 652 Rhomboplites aurorubenes, 707 Rhombosolea tapirina, 455 Rhyacodrilinae, 205 Rhynchactis, 474

1402

Oceanography and marine biology Rhynchohyalus, 474, 488 Roboralga jacksoniensis, 707 Rotifera, 727 Roughleyia australis, 707 Saccharomyces carlsbergensis, 559 ” cerevisiae, 610 Sagitta, 648 ” elegans, 646 ” hispida, 726 Salarias fasciatus, 435 Salicornia virginica, 657 Salmo clortii, 707 ” gairdneri, 546, 706 ” malma, 706 ” salar, 605, 737 Salmoniformes, 474, 488 Salpa fusiformis, 648 Salvelinus fontenalis, 551, 569 Saradinia caerlea, 708 Sarcodia, 652 Sordina pichardus, 706 Sardinella aurita, 708 ” fimbriata, 708 Sargassum, 656 ” grevillei, 656 ” horneri, 648, 656 ” kjellmanianum, 656 ” ringgoldianum, 656 ” sagamianum, 656 ” tenerrimum, 656 ” thunbergii, 656 Sarpa, 459 ” salpa, 457, 459 Saurida undosquamis, 708 Saxostrea commercialis, 680 Scaphander lignarius, 680 Scartelaos histophorus, 434 Scenedesmus, 548, 575 ” obliquites, 719 ” quadricauda, 723 Schizodesma, 362 Sciaena antarctica, 708 ” coitor, 708 ” maculata, 708 Scillium caniculata, 708 Scirpus paludusus, 657 ” validus, 657

929

Interlinking of physical

1404

Scomber australasicus, 708 ” scrombus, 708 Scomberomoms, 708 ” maculatus, 708 Scopelarchus analis, 526 Scophthalmus aquosus, 709 ” rhombus, 454 Scopimera intermedia, 686 Scorpaena guttata, 709 Scrippsiella faeroense, 723 Scrobicularia, 671 ” plana, 563, 575, 602, 658, 670 Scutellidium, 732 Scylla serrata, 574, 613 Sebastes auriculatus, 709 ” nivosus, 708 ” paucispinis, 708 ” thompsoni, 708 Secutor, 474, 485 Selar, 709 Selaroides leptolepis, 709 Selenastrum, 542 ” capricornutum, 541, 554, 618 ” minutum, 719 Semibalanus balanoides, 574 Sepia, 674 ” officinalis, 536, 570, 673 ” teuthris indica Sepioteuthis indica, 675 Sergestes similis, 525 Seriolia grandis, 709 Serpulidae, 266 Sicyases, 433 ” sanguineus, 433, 434, 435, 457 461 ” ingentis, 684, 686 ” oramin, 708 ” rivulatus, 708 Sinonovacula contricta, 672 Siphamia, 474, 485, 491 Skeletonema costatum, 101, 541, 548, 554, 555, 606, 620, 723 Solaster papposus, 688 Solea solea, 710 Soleidae, 455 Solenostoma crenulatum, 548 Spartina, 243, 586, 605, 611, 657, 715 ” alterniflora, 561, 579, 627, 657, 714 ” patens, 657 Spathoglossum asperum, 656

Oceanography and marine biology Spinachia spinachia, 457, 460 Spiniphryne, 475 Spiridion, 204 ” insigne, 212, 214, 215, 247, 286, 287, 294 Spirorbis lamellosa, 725 Spisula solidissima, 571, 672 Spondylus americanus, 672 Sporozoa, 246 Sphagemacrurus, 474, 487 Sphyraena novae-hollandiae, 710 ” sphyraena, 710 Sphyrna lewini, 710 ” tiburo, 710 Sprattus sprattus, 710 Squalogadus, 488 Squalus acanthius, 710 ” mitsukurii, 710, 712 Squilla oratorio, 686 Staphylinidae, 244 Staphylococcus epidermidis, 620 Steindachneria, 474 ” argentea, 487 Stellifer lanceolatus, 711 Stenobrachius, 496 Stercutus niveus, 218 Sternoptyx, 497 Stichaeidae, 440, 443 Stoechospermum marginatum, 656 Stomias, 511 Streptococcus faecalis, 109 ” lactis, 100 Strombus gigas, 680 Strongylocentrotus droebachiensis, 688 ” franciscanus, 688 ” purpuratus, 688 Stylaria lacustris, 293 Symbolophorus californiensis, 523 Symphurus plagiusa, 711 Symplectoteuthis oualaniensis, 675 Synaptura nigra, 711 Syringodium filiforme, 658 Taaningichthys paurolychnus, 508 Tadorna tadorna, 244 Tapes decussatus, 563, 672 ” japonica, 672 Taphius glabratus, 672 Tarletonbeania, 525

931

Interlinking of physical ” crenularis, 523, 525 Tarpon atlanticus, 711 Taurulus bubalis, 437 Tautoga onitis, 711 Tautogolabrus adspersus, 538 Tealia felina, 660 Tegula funebralis, 549, 680 ” galina, 679 ” viridula, 679 Tellina, 713 Tellina tenuis, 544, 713, 718, 735 Telosporidia, 246 Terebra, 351, 355 ” salleana, 354 Tetraselmis, 723 Thais emarginata, 680 ” lapillus, 578, 679 Thalassia, 577, 658 ” testudinum, 576, 586, 590, 657 Thalassiosira fluviatilis, 723 ” pseudonana, 548, 554, 555, 565, 619, 718, 723 ” weissflogii, 602 Thalassodriloides belli, 265, 266 Thalassodrilus, 204 Thaumatichthyidae, 475 Thaumatichthys, 475 Thioploca, 116 Thunnus albacares, 711 ” thynnus, 710 Tigriopus californicus, 461 ” japonicus, 565 Tilapia, 547, 569, 740 ” musambica, 738 Tisbe, 565 Tivela crassatelloides, 680 Tomicodon boehlkei, 429, 443, 444, 452 ” eos, 443, 444, 448, 452 ” humeralis, 429, 436, 443, 454 ” myersi, 443, 444, 448, 452 ” zebra, 442, 444, 452 Torula (=Candida) utilis, 559 ” utilis, 720 Trachichthodes affinis, 711 Trachichthyidae, 474, 478, 483 Trachonurus, 474 Trachurus trachurus, 711 Trachypencus constrictus, 686 Tricladida, 245

1406

Oceanography and marine biology

933

Trigla lucerna, 711 Triglochin maritima, 658 Trinectes maculatus, 455 Tringa totanus, 244 Tripterophycis, 474, 488 Tripterygiidae, 442, 454 Tripterygion, 439, 451 ” capito, 454 ” robustum, 440 Trisopterus luscus, 242, 711 Tritia, 375 Tubellaria, 229 Tubifex, 205, 243, 317, 324, 325, 591 ” costatus, 201, 204, 206, 207, 212, 213, 214, 221, 222, 223, 225, 226, 228, 229, 230, 231, 233, 238, 242, 258, 260, 265, 268, 269, 276, 277, 279, 282, 287, 289, 291, 292, 293, 296, 297, 300, 301, 305, 320, 321, 326 ” postcapillatus, 265 ” pseudogaster, 201, 204, 206, 207, 212, 213, 214, 221, 222, 226, 228, 229, 231, 233, 242, 244, 258, 265, 277, 292, 293 ” tubifex, 212, 229, 230, 244, 279, 281, 293, 296, 298, 300, 302, 305, 312, 313, 314, 316, 318, 319, 320, 322, 323, 324 Tubificidae, 198, 199, 204, 206, 212, 225, 227, 229, 277, 278 Tubificinae, 205 Tubificoides, 205, 238, 287 ” aculeatus, 287 ” apectinatus, 206, 207 ” gabriellae, 206, 207, 226, 229, 242, 282, 290, 298, 300, 308, 322 ” heterochaetus, 206, 207, 293, 321, 326 ” intermedius, 263 ” longipenis, 263, 264, 290 ” multisetosus, 239, 322 ” nerthoides, 206, 207 Turbellaria, 244, 245 Turbo cornutus, 681 ” smaragdus, 679 Tylos punctatus, 687 Typhlogobius californiensis, 437 Typhloplanoida, 245 Udotea flabellum, 646 Ulothrix, 542 Ulva, 110, 239, 648, 650 ” fasiciata, 649 ” lactuca, 649 ” pertusa, 649 ” reticulata, 649 Umbrina roncadar, 711 Undinula vulgaris, 732

Interlinking of physical Uniporodrilus, 198, 204 Upeneus moluccensis, 712 Upogebia pugettensis, 687 Uronema marinum, 561 Urophycis chuss, 712 Urosalpinx, 403 ” cinerea, 563 Uteriporus vulgaris, 245 Valenciennellus, 518 Valonia fastigiata, 650 Valoniopsis pachynema, 650 Vargula (=Cypridina) hilgendorfi, 492 ” hilgendorfi, 492 ” tsujii, 495, 496 Venerupis decussata, 735 ” philippinarum, 671 Ventrifossa, 474 Venus gallina, 672 ” verrucosa, 671 Vibrio, 475 ” albensis, 476 ” anguillarum, 546 ” fischerei, 475 ” harveyi, 475 ” logei, 475 ” splendidus, 475 Villorita cyprinoides var. cochinensis, 563, 735 Vinciguerria attenuata, 712 Wapsa, 205 Watersipora cucullata, 726 Winteria, 474, 488 Xenodermichthys copei, 519 Xererpes fucorum, 424, 433, 443, 445, 447, 450 Xiphigorgonia anaps, 660 Xiphister atropurpureus, 428, 433, 437, 441 ” mucosum, 440 ” mucosus, 432, 435, 436, 437 Xiphopeneus kroyeri, 687 Yarrella, 497 Zeus australia, 711 Zoarces viviparus, 712 Zostera, 235, 237, 239, 284, 358, 657

1408

Oceanography and marine biology ” marina, 561, 579, 657 ” muelleri, 657

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Interlinking of physical

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SUBJECT INDEX References to complete articles are given in heavy type; references to sections of articles are given in italics; references to pages are given in normal type.

Aanderaa current meters, 64 Abalone, 553 Adelie penguin, 5, 6, 7, 8, Fig. 9 (Tranter) facing p. 8 Adrenaline, 505 Adrenocorticotropin (ACTH), 429 Adriatic Sea, copper in fishes, 566 ” intertidal fishes, 424 Aegean Sea, copper, 637 Africa, intertidal fishes, 420 ” upwelling, bacteria, 97 Aggression of intertidal fishes, 452–3 Agnes, tropical storm, 597, 659 Alanine as end-product to anaerobiosis in Bullia, 396 Alaska, copper, 595, 637 ” intertidal fishes, 422 Algae effect of copper, 559–61, 648 Alginates, 583 Allantoinase in polychaetes, 718 Alpha Helix South East Asia Bioluminescence Expedition, 481 Altamaha River, copper, 629 Alvin accident, bacteria, 110, 111 American Chemical Society Symposium, 538 Amylase, 718 Añasco River, Puerto Rico, copper in delta, 715 Anchovy, 547 Angler fish, 488 Anodic stripping voltammetry (ASV) to differentiate between strongly and weakly complexed copper, 606 Anoxic environments, methane production, 105–6 Antarctic and North America, birds, 569 ” ” petrels, 714 ” benthic species, 5 ” bottom water, 12, 136 ” Convergence, 17, 20, 23, 29, 30, 31 ” ” bacteria, 101 ” day-length, 19 ” Divergence, 9, 14, 29, 31

Interlinking of physical

1412

” ecosystem, 2–5; 1, 26, 31 ” ” birth of, 14 ” habitat, 28–31 ” ice community, 1 ” Intermediate Water, 12 ” krill, 5, 6, 8, 27, 29, 30, 31, 32, Fig. 4 (Tranter) facing p. 4 ” light, 19–26 ” limits of growth, 6–9 Antarctic meridional circulation, 15 ” nutrients, 9–19 ” Ocean, interlinking of physical and biological processes in, 1–35 ” pack ice, 2 ” ” system, 26–7 ” plankton productivity, 1 ” Surface Water, 12, 29 ” upwelling, 26 ” ” areas and copper, 538 Antibiotics and bacteria, 96 Arabinose, 602 Aral-Caspian basin, copper, 629 Arctic Ocean, 2, 28 Asia, copper, 629 Atlantic, authochthonous organic matter, 83 ” bacteria, 99 ” in surface film, 107 ” blennies, 426 ” copper, 629, 634, 637, 638, 642 ” grey seal, 550, 613 ” Ocean, 38, 40, 45, 46, 47, 57, 59 ” particulate matter, 136, 138, 141, 142, 151 ” salmon, 567, 606 ” sandy beach whelks, 370 ” silverside, 546 ” trace metals, 586 Australasia, intertidal fishes, 420 Australia and New Zealand coastal copper, 643 ” copper, 594 ” pollution, 567, 576 Australian estuary, bacteria, 96, 97 Azores, intertidal fishes, 422 Azov Sea, copper, 629 Bacteria, adenosine triphosphate (ATP), 87, 114 ” Alvin accident, 110, 112 ” amino acids, 86, 90, 91, 93, 94, 95 ” and copper, 644 ” diagenesis in marine sediments, 89 ” primary production, 96–8

Oceanography and marine biology

939

” attached to particles, 100 ” attachment to organic particles and detritus, 87 ” barotolerance, 109, 110, 112 ” C:N:P ratios, 96 ” distribution in the sea, 106–12 Bacteria, effect of copper, 557–9 ” epifluorescence microscopy, 97 ” grazing of, 100–6 ” heterotrophy with nitrate and sulphate as terminal electron acceptor, 103–5 ” in deep-water columns, 109–12 ” in deep-sea sediments, 112–5 ” in surface film of sea water, 107–9 ” rôle of in turnover of organic matter in the sea, 74–127 ” trophic rôle, 98–100 ” turnover rates of organic matter, 91– 4 Bacterial, activity, effect of starvation, 100–1 ” inorganic chemical composi-tion of the sea, 102–6 ” biomass, chemical estimates, 88–90 ” estimation of, 85–9 ” by autora-diogra-phy, 86 ” by culture tech-niques, 86–7 ” by microscopy, 85–6 ” light organs in Apogonidae, 484– 5 ” Anomalopidae, 478–83 ” Bathylagidae, 488 ” Beryciformes, 478– 84 ” Chlorophthalmidae, 488 ” fish, distribution of, 477–89 ” Gadiformes, 487– 8 ” Leiognathidae, 486 ” Lophiformds, 488– 9 ” Macrouridae, 487 ” Merlucciidae, 487 ” Monocentridae, 484 ” Moridae, 488 ” Perciformes, 484– 6 ” Salmoniformes, 488 ” Trachichthyidae, 483 ” luciferase, 476, 492, 493, 494, 495, 497 Bacterial, metabolism, 75, 76, 78 ” production, 90–8 ” 14C tracer techniques, 90–1 ” respiration, 91, 95, 98, 104, 111, 113 ” symbionts of fishes, 473–4 Bacteriostatic compounds in certain waters, 101 Bahamas, flux of faecal pellets, 81 ” intertidal fishes, 421 Baleen whales, 1, 5, 6, 7, 30, 32

Interlinking of physical

1414

Baltic Sea, copper, 629, 634, 638 ” from Sweden, 637 ” metal concentration, 637 ” sediments and heavy metals, 608 Barents Sea, copper, 634 Batfish, 496 Bathysphere, Beebe, 526 Beaufort Sea (Mackenzie River Delta) copper, 643 BELGICA, 1 Benthic organisms and copper, 648–712 Bering Sea, copper, 629 Bioassay techniques for testing water quality, 553 Biogenic particles, settling velocities, 164 Biogeochemistry of copper in sea, discussion of, 628 Biological importance of copper in oceans and estuaries, 535–789 ” Investigations of Marine Antarctic Systems and Stocks (BIOMASS), 1, 32 Biology and ecology of marine Oligochaeta, 197–350 Biology of intertidal fishes, recent studies, 420–71 ” sandy-beach whelks of genus Bullia, 351–418 Bioluminescence in fireflies, 527 ” tunicates, 489 ” fishes, adrenaline as a neurotrans-mitter, 505 ” adrenaline injection, 499, 503, 505, 506, 511, 512, 514 ” adrenergic control, 508 ” alpha-adrenergic agonists, 505 ” alpha-adrenocep-tor antagonists, 507 ” amphetamine injections, 503 ” and predation, 481 ” aspects of 473– 533 Bioluminescence in fishes, Atlantic, 493 ” beta-adrenergic agonists, 505 ” “blink and run” escape behaviour, 481 ” blinking rates, 480 ” California, 494, 495, 496 ” Caribbean, 480 ” catecholaminergic endings destroyed by 6-hydrocydopamine, 506 ” counter-illumination, 515, 518, 519, 522, 523, 525, 526 ” dopaminergic nerve endings destroyed by 6hydroxydopamine, 506 ” effect of adrenaline, 508 ” effect of noradrenaline, 514 ” electrical responses of photophores, 503, 505, 507, 514 ” electron microscope studies, 474, 486, 488, 489, 505 ” 5 HT as a neuromediator, 505 ” Indonesia, 480 ” Indo-Pacific, 480 ” induced by 6-hydroxydopamine, 506 ” induced by K+, 506 ” light emission and oxygen consumption, 503, 505

Oceanography and marine biology

941

” neurotransmitter control, 506 ” noradrenaline injection, 495, 503, 505, 506, 507, 511, 514 ” noradrenergic nerve-endings destroyed by 6hydroxydopamine, 506 Bioluminescence in fishes, North America, 494, 499 ” pharmacological effects, 505, 515 ” pharmacological responses of photophores, 503, 505, 507 ” Philippines, 494 ” Puget Sound, 494, 495, 496 ” Red Sea, 480 ” regulation of, 523 ” response to adrenaline, effect of 5-hydroxytryptamine, 505, 506 ” response to adrenaline, reduced by dopamine, 505 ” responses to stimulation of isolated photophores, 512 ” seen from a submersible, 487, 526 ” stimulated by adrenaline, 487, 499, 501, 507, 511, 512 ” stimulated by isoproterenol, 512 ” stimulated by noradrenaline, 511, 512 ” stimulated by phenylphrine, 511 ” study of larvae to determine when light organs become effective, 491 ” ventral photophores, 518, 519, 523 ” yellow lenses, 526 ” rôle in marine environment, 527 Bioturbation, 84, 85 Black marlin, 567 ” River, copper, 629 ” Sea, bacteria, 102, 103, 104, 114 ” copper, 629, 642 ” trace metals, 743 Blennies, 431, 439, 446, 452, 456, 457, 461 Bluefish, 575 Bluegill, 569 Bornholm Sea, copper, 638 Britannia Beach, B.C., copper from mining, 625 British Columbia, intertidal fishes, 422, 424, 427, 443 Brook trout, 549 Bullia, accumulation of metals, 407 ” A.T.P. concentration, 397 ” adenylate energy charge (AEC), 408; 397 ” an activity budget, 391–2 ” anaerobiosis, 396–7 ” and ammonium nitrate pollution, 403 ” pollution, 403–8 ” water loss, 372 ” behavioural and pathological effects of pollutants, 404–5 ” biology of, 351–418 ” blood circulation, fluid spaces, and foot extension, 376–82 ” system, 376–80 ” volumes, 381

Interlinking of physical

1416

” burrowing, 386–8 ” chemoreception, 399–400 ” circumenteric nerve ring, 400 ” contact behaviour, 358 ” copulation, 374 ” cost of free existence, 393–5 ” crawling, 385 ” digging activity, 386, 388 ” distribution on beach, 357–8 ” effect of oil, 405–6 ” oil dispersants, 404, 405 ” pollution on oxygen uptake, 406–7 ” salinity on oxygen uptake, 394–6 ” egg capsules, 374–6 ” elimination of foreign particles, 409 ” emergence from sand, 388 ” energy costs of activities, 390–1 ” energy expenditure, 391 ” environmental tolerance limits, 370– 2 ” food and feeding, 358–66 ” foot extension, 380 ” general ecology and behaviour, 355– 71 ” genetics, 410 ” haemocytes, 409 ” haemolymph, 409 ” India, 375, 376, 388 ” larval suppression, 376 ” values, 371 ” locomotion, 382–8 ” mechanoreception, 400 ” nervous system, 402–4 ” oxygen consumption, 388–90; 392, 393, 404 ” parasites, 409 ” parasitology and pathology, 409 Bullia, population studies, biomass and chemical composition, 368–70 ” predators of, 366–7 ” production/biomass ratios, 368, 369 ” reproduction, 373–6 ” reproductive cycles, 373–4 ” respiration, metabolism, and energy utilization, 388–97 ” response to amino acids, 400 ” γ-buterobetaine, 400 ” indole, 400 ” skatole, 400 ” trimethylamine, 400 ” salinity tolerance, 371–2 ” sea-water spaces, 381–2 ” sensory systems, 399–403

Oceanography and marine biology

943

” South Africa, 366, 376, 388 ” streptoneury, 402 ” surfing, 382–6 ” surfing habit and response currents, 354 ” temperature tolerance, 370–1 ” tidal migrations, 355–7 Cadmium, effect on Bullia, 404, 405 California, copper, 587 ” intertidal fishes, 422, 424, 427, 430 ” metals in waste water, 586 ” pollution, 563, 567, 631, 638 Californian grunnion, 438 ” harbours, copper, 622 Cape Leeuwin, Australia, ferromanganese nodules, 623 ” St. Vincent, Mediterranean water, 38, 40, 41, 45, 47, 48, 51, 59, 60, 63, 72 ” Santa Maria, 64 Cariaco Trench, anoxic waters, 189 ” bacteria, 104, 106 Caribbean Sea, bacteria in surface film, 107 ” particulate matter, 136 Carnivore efficiencies, 79 Carp, 569 Catfish, 603 Cenozoic sediments near crest of Mid-Atlantic Ridge, 625 Ceylon, intertidal fish, 448 CHALLENGER, 1, 41 Challenger data, 53, 70 ” Expedition, 625 ” survey, 50 Channel catfish, 546 ” Islands, intertidal fishes, 422 Chelation, defined, 607 Chemical speciation of organic compounds in the sea, 76–7 Chemoautotrophic bacteria, 114, 115 Chesapeake Bay, copper, 563, 597, 637 ” dumpsite, copper, 630 ” sediments, copper, 630 Chilean clingfish, 458 China Sea, copper, 638 Circumpolar circulation, 14, 29 ” Current, 14, 29 Clingfishes, 429, 431, 434, 443, 444, 448, 452, 456, 457, 461 ” associated with sea urchin, 444, 449 Coccoliths, 171 Coefficient of biological absorption (CBA), 569 Coelenterazine, a coelenterate-type luciferin, 496, 497, 498 Coho salmon, 545, 546, 557, 589

Interlinking of physical

1418

Columbia River, copper, 594, 638 Conductivity-temperature-depth (CTD) casts, 64 Conductivity-temperature-depth (CTD) casts, Mediterranean outflow, 57 Conductivity-temperature-depth (CTD) casts, observations, 66 Connecticut River, Long Is. Sound, copper, 604, 630 Controlled Ecosystem Pollution Experiments (CEPEX), 542, 557, 558, 566, 568, 713, 742 Convention for the Conservation of Antarctic Living Resources, 32 Conway estuary, U.K., copper, 641 Cooper River, copper, 629 Copepod faecal pellets, 82 Copper “acute” effects, 715 ” adaptation of organisms to elevated levels, 547, 548 ” adsorption by hydrous alumino-sili-cate clays, 617 ” algae, 560–1, 648; 649, 651, 653, 654, 656, 657 ” and amylase and trypsin activity, 539 ” cadmium, biological effects, 619– 20 ” cobalt, biological importance, 621 ” depression of photosynthesis, 540, 541 ” dredging, 626–8 ” iron, biological importance, 620 ” lead, biological effects, 619 ” magnesium, biological importance, 621 ” manganese, biological importance, 621 ” marine mammals and birds, 713 ” mercury antifouling paints, 740 ” mercury, biological effects, 619 ” nektonic organisms, 689–712 ” nickel, biological effects, 619 ” other metals, interaction with organics as a method of control, 550 ” oxygen consumption, 545 ” production of chlorophyll, 537 Copper and production of huemes and incorporation into haemoglobin, 536 ” sperm activity, 537 ” zinc, biological effects, 619 ” antifouling compounds, 621–3 ” behavioural effects, 543 ” biological availability controlled by complexing agents, 602, 603 ” change in effect during life of an organism, 738–41; 546 ” -chelation capacity of sea water measured by bacteria, 553 ” “chronic” effects, 715 ” -complexing agents, 606, 607 ” -containing herbicides, cutrine, 542 ” metalloenzymes, 537 ” prosthetic group in haemocyanin, 536 ” dissolved, 601–14 effects on bacteria, 557–9

Oceanography and marine biology ” bioassay organisms, 554– 7 ” Bullia, 405 ” fishes, 544, 545, 546, 547 ” flux of nitrogen through planktonic food chain, 541 ” fungi, 559 ” growth, 540 ” invertebrates, 562–6 ” larval developments, 566, 741 ” organisms, 536–70 ” beneficial, 536– 8 ” detrimental, 539–53 ” osmoregulation, 545 ” oxygen consumption, 545 ” protein, 541 ” Protozoa, 561 ” RNA, 541 ” sea urchin spermatozoa, 573 ” vertebrates, 566–9 ” essential for growth and metamorphosis, 537 ” fate in estuaries, 589 ” fluxes into marine environment, 629, 630, 632 ” from mine waste site, 548 ” from nuclear and fossil fuel power plants, 589 ” granules, 549 ” in parenchymatous tissue, 545 ” histological effects, 543 Copper in Antarctic upwelling areas, 538 ” bacteria, 645 ” benthic organisms, 648–712 ” biocidal agents, 623 ” haemocyanin, 536 ” haemolymph, 537 ” invertebrates, 649–89 ” marine phytoplankton, 644–5 ” marine zooplankton, 645–8 ” oceans and estuaries, 535–789 ” planktonic organisms, 644–8 ” inhibition of fertilization, 566, 739 ” levels at air-sea interface, 632 ” found to be toxic to marine organisms, 715–38 ” in interstitial (pore) water, 641 ” marine environment, 628– 42 ” organisms, 642–715 ” sediments, 640–2; 643 ” waters, 634–40 ” metal-metal interactions having biological importance, 618–20 ” metal species and biological effects, 620 ” models used to determine fate in marine environment, 551

945

Interlinking of physical

1420

” organs accumulating and storing it, 549, 550 ” particulate and adsorption, 613–6 ” physiological effects, 543 ” poisoning of larval enzymes, 510 ” prediction of biological effect, 550 ” source from tubing in desalination plants and power stations, 552, 553 ” sources of, 535 ” and changes of potential biological importance in copper species in marine environment, 584 ” transfer through food chain, 712–5 ” uptake and accumulation by organisms, 570–84 ” water quality, criteria, what to measure, 742–3 Coriolis factor, 58 ” force, 39 Crabeater seal, 5, 6, 7, 32 Cunner, 538 Current velocities in Mediterranean outflow, 61–72 Cuttlefish, 537 Cytochrome oxidase, 537 Dead Sea sediments, trace elements, 616 Deep-ocean mining, 538 Deep-sea bacteria, isolation of, 111 ” fauna, rôle in organic degradation, 111–2 ” metal deposits, 623–6 Deep-sea sediments, amino acids, 76 ” bacteria in, 112–5 ” carbohydrates, 76 Delaware Bay, copper, 597 Denmark Strait overflow, 137 Department Fishery Data Centre, FAO, 689 Desalination plant, Gulf of Elat, 589 ” Key West, Florida, 552 ” plants, 562 Diablo Canyon nuclear power plant, California, 552, 553 4, 6 diamidino-2-phenyl-indole (DAPI), 86 DNA stained by acridine orange, 86 DISCOVERY II, 1, 38, 41 Discovery hydrographic sections, 40 Dissolved organic carbon (DOC) in the sea, 76; 76, 77, 78, 79, 81, 84, 87, 91, 96, 97, 98, 99, 100, 104, 107, 108, 115 ” production by phytoplankton, 77 ” matter, marine biologists’ definition, 75 Distribution of free-living luminous bacteria, 476 Dolphin, 570 Dover sole, 545, 567 Dredging, effect on metal content of sediments, 627 Du Pont dumpsite, U.S.A., copper, 630 Dutch coast, intertidal fishes, 422, 427

Oceanography and marine biology

947

” Wadden Sea, copper, 634 ” intertidal fishes, 424, 445 ” metals in interstitial water, 615 Dyking areas, effect of removal of tides on exchange of interstitial water, 615 East Equatorial Pacific Ocean, 136 ” Pacific Rise, 625 ” Wind Drift, 9, 29 Ecosystem of Antarctic, 2–5 Eels, 546 El Niño and anchovy fishery, 104 ELTANIN, 1 England, intertidal fishes, 431 English Channel, copper, 634, 637 Epibacteria, 87 Epifluorescence microscopy, 86 Equatorial Atlantic Ocean, copper, 635 Ergotamine, alpha-adrenergic agonist, 505 Europe, intertidal fishes, 420, 422, 428, 430 European Atlantic coast including North and Baltic Seas, copper, 642 ” Mediterranean coast, copper, 638 Evolution of luminous bacterial symbioses, 489, 490, 493 Fatty acids, indicators of bacterial biomass, 87, 88, 89 Ferromanganese nodules, 623, 624 Finfish, 567 First International Biomass Experiment (FIBEX), 32 Fishes, bacterial symbionts, 473–4 ” bioluminescence, 473–533 ” , counter-illumina-tion, 515–27 ” distribution of bacterial light organs, 477–89 ” photophore distribution and ambient light, 515–8 ” physical correlates of ventral bioluminescence, 519–27 Fladen ground, North Sea, 78 Flounder, 718 Follicle stimulating hormone (FSH), 429 Fraser River, copper, 594, 597 FRAM, 32 France, intertidal fishes, 422, 443 “Fulvic” compounds and biological availability of copper, 607, 608 Fungi, effect of copper, 559 Fur seal, 5, 7 GAUSS, 1 Geochemical Oceans Sections Study, 129 Geochemistry of natural series radionuclides, 179 GEOSECS, 137, 142 German Atlantic Expedition, 38

Interlinking of physical

1422

Glutamic oxaloacetate transaminase, 564 Glycoprotein as an antifreeze, 8, 11 Gobies, 425, 431, 442, 443, 452, 455, 459, 460, 462 Golden shiner, 546 Goldfish, 567, 603 Great Britain coastal copper, 638, 641, 642 ” intertidal fishes, 422 Gulf grunnion, 438 ” of Cadiz, Mediterranean water, 38, 39, 40, 41, 44, 50, 53, 59, 61, 63, 64, 66 ” Mediterranean water outflow in 38–73 ” California, intertidal fishes, 421, 422, 423, 424, 429, 431, 432, 442, 443, 444 ” Elat, desalination plant, 589 ” Gdansk, copper, 593 ” Maine, copper, 637 ” Mexico, copper, 588, 638, 643 ” particulate matter, 136, 151 ” St. Lawrence, copper, 587, 594 Hatchetfish, 497, 518, 526 Herbivore grazing efficiencies, 78, 79 Herring, 553 Holland, acute fish toxicity, 585 Homing and orientation of intertidal fishes, 450–1 Horseshoe crab, 537 Housatonic River, Long Is. Sound, copper, 630 Howe, Sound, B.C., copper, 641 Humber estuary, U.K., copper, 588 “Humic” compounds and biological availability of copper, 607, 608 Hydrothermal vents, bacteria, 106 ” in deep sea, 114–5 6 hydroxydopamine (6-OHDA), 506 “Ice” algae, 26, 27, 28 India, copper in Cochin backwaters, 562 ” sandy-beach whelks, 352, 362, 366, 377 Indian Ocean, 12, 21 ” copper, 637, 638 ” nodules, 623 ” particulate matter, 136, 151 ” rates of metal accumulation in sediments, 623 Innervation of photophores and chromatophores, diagram, 499 Institute of Oceanographic Sciences, 163 International Copper Research Association Inc., 536 ” Geographical Year, 1 ” Indian Ocean Expedition, 9 ” Symposia on Environmental Biogeochemistry, 539 Intertidal fishes, adaptation to the physical environment, 426–35 ” Adriatic Sea, 458

Oceanography and marine biology ” Barbados, 458 ” British Columbia, 450, 460 ” California, 449, 452, 459 ” Caribbean, 453 ” Ceylon, 449 ” Chile, 458 ” community and population surveys, 421–4 ” damaged by sea urchin, 432 ” definition and classifications, 421 ” development, growth and longevity, 441–2 ” Dutch Wadden Sea, 460 ” egg cleaning, 440 ” emersion, 432–5 ” endogeneous tidal rhythmicity, 454 ” Europe, 449 ” extent of movement, 447– 50 ” fidelity to specific pools, 449, 450 ” food and feeding, 456–60 ” Gulf of California, 450, 457 ” habitats, 427 ” Hawaii, 452 ” homing and orientation, 450– 1 Intertidal fishes, intertidal movements, 447– 56 ” Japan, 460 ” Madagascar, 452 ” Mediterranean, 456’ ” memory, 451 ” methods of collecting, 422, 423 ” morphology, 425–6 ” pheromones, 438, 439 ” population densities, 424 ” predation, 461 ” prolactin activity, 429 ” recent studies on biology of, 420–71 ” respiration, 435–6 ” reproduction, 437–41 ” rhythmic behaviour, 453– 6 ” salinity, 427–9 ” “satellite” males, 439 ” Scotland, 459 ” South Africa, 450, 459 ” stomach contents of pigeon guillemot, 461 ” “summer kill”, 431 ” temperature, 429–31 ” territoriality and aggression, 452–3 ” turbulence and wave action, 431–2 ” visual pigments, 426 ” “winter kill”, 431

949

Interlinking of physical

1424

” zonation and habitat selection, 442–7 ” invertebrates as indicators of heavy metal pollution, 555 Invertebrates and copper, 561–6, 648–89 Irish Sea, copper, 635 Iron and humic compounds as complexive agents in polluted water, 542, 543 Isoleucine, 602 Isoproterenol, beta-adrenergic agonist, 505 Japan, intertidal fishes, 424 ” Sea, copper, 629 Japanese puffer, 438 JEAN CHARCOT, 38, 39 Jean Charcot cruise, 42 J.F.Kennedy Space Center, Florida, 595 Kamchatka peninsula, intertidal fishes, 422 Kaohsiung Harbor, Taiwan, copper, 587 Kara Sea, copper, 629 Kattegat, trace metals, 743 Kerguelen-Gaussberg Ridge, 26, 29 Key West, Florida, dumpsite, copper, 630 Killifish, 568 King salmon, 568 Korea, intertidal fishes, 424 Kuril Islands, intertidal species, 442 Labrador Sea Water, 136 Lagoon of Grado, Italy, copper, 638 ” Venice, Italy, copper, 638 Lamont-Doherty Geographical Observatory, 134 ” nephelometer, 134 Lanternfishes, 515, 522, 523, 526 Laptev Sea, copper, 629 Lead, effect on Bullia, 405 Light, Antarctic, 27–6 Light in the sea, angular and spectral distribution of, 494 Ligurian Sea, zinc cycle, 579 “Littoral coastal species”, 426 “Littoral fishes”, 420, 421 Lipid phosphate, indicator of bacterial biomass, 87 Long Is. Sound, copper, 87 Los Angeles, copper, 631 ” harbour, dredging, 627 Luciferace, 476, 483, 492 Luciferase, 476, 489, 492, 493, 495, 497, 498 ” extracts from Euphausia, 494, 495 ” in Diaphus, 496 Luciferin, 485, 493, 494, 495, 496, 497, 498, 507

Oceanography and marine biology

951

” chemical nature of, 496 ” Cypridina, 492, 493, 494, 495, 496, 497 ” crustaceans, 495 ” ostracods, 485 ” Renilla, 496 Luciferin-luciferase extracts from firefly, 492 ” reactions, 492, 497 ” system, 497 ” in Cypridina, 492 Luminescence in copepods, 497 Luminescent system, dietary requirements, 492 Luminous bacteria, 474, 475, 476, 477, 481, 483, 484, 485, 487, 488, 489, 491 ” and temperature, 476 ” Atlantic, 477 ” autoinduction, 476 ” California, 477 ” catabolite repression, 476 ” chemical systems, 492– 8 ” distribution of free-living, 477 ” electron micrographs, 483, 486, 488 ” general considerations, 489–92 ” Gulf of Elat, 477 ” kinetics of, 477 ” Mediterranean, 477 ” oxygen equipment, 476 Luminous bacteria, physiology, 476 ” polyhydroxybutyrate (PHB) storage granules, 483, 486, 488 ” properties of, 475–6 ” systematics of, 475–6 ” Texas, 477 ” responses and correlated structural features of photophores of Porichthys, diagram, 508 Luteinizing hormone (FH), 429 Lysine oxidase, 537 Mackenzie Valley, Canada, sediments, effect on Beaufort Sea water, 592 Madagascar, intertidal fishes, 422 Malaysia, intertidal fishes, 422 Manganese nodules, 623, 624, 625 ” mining and dredging, 621–8 Mangrove swamps, intertidal fishes, 422 ” uptake of metals, 580 Marine bacteria, copper, 720 ” taxonomy of, 85 ” biogeochemical cycle of copper, 535 ” birds, 569 ” chemoautotrophic bacteria, 102–3 ” fungus, copper, 720 ” mammals, 570

Interlinking of physical ” and birds, copper, 712 ” oligochaete groups, 198–200 ” organisms, copper levels in, 642–715 ” toxic copper levels, 715– 38 ” sediments, copper levels, 640–2 ” snow, 84, 148, 168 ” sulphate-reducing bacteria, copper, 720 ” waters, copper levels, 634–40 ” yeast, copper, 720 ” inhibition by copper, 738 Mediterranean, Atlantic inflow, 57 ” blennies, 426, 429, 460 ” copper, 637, 642 ” deep countercurrent, 72 ” intertidal fishes, 422, 432, 443 ” Israel, copper, 638 ” metal levels in mussels, 563 ” outflow, 1970–1980, 40–1 ” Challenger cruise 1976, 50 ” current velocities, 61–72 ” currents, 59, 61, 64, 65, 66, 67, 69, 70, 72 ” formation of tem-perature-salinity maxima, 50–8 Mediterranean outflow, hydrographic survey, 41–50 ” in Gulf of Cadiz, 38– 73 ” mixing in, 58–61 ” schematic section, 55 ” Shackleton cruise, 1973, 41–50 ” temperature, 66 ” transport, 61, 63, 64 ” sandy-beach whelks, 370 ” Sea paniculate organic matter, 76, 149 ” phytoplankton, 541 ” pollution, 565, 569 ” undercurrent, 41, 47, 50, 58, 61, 64, 69, 72 ” water, 40, 44, 59, 60, 69, 72 Melanin in Bullia, 399 Melbourne, Australia, cadmium pollution, 573 Melt-water, spreading of, Fig. 16 (Tranter) facing p. 15 Menke’s syndrome, 537 Mercury effect on Bullia, 405 Metal deposits, deep-sea, 623–6 Metals altering membrane permeability, 539, 540 ” as enzyme inhibitors, 539, 540 ” distortion of nuclei acid structure, 539 ” impairing protein synthesis, 539 ” inhibiting phosphorylation, 539 Metalloenzymes, 546 Metallo-organic complexes, influence on trace

1426

Oceanography and marine biology

953

metals, 601 Metalloproteins, 537, 716 Metallothionein, 537, 550, 567, 574, 612, 620 ” -like complexing agent, 582 ” proteins, 574, 585, 610 METEOR, 41 Meteor Expedition, 64 ” data, 66 Methane production in anoxic environment, 105–6 Methanogenic bacteria, 105, 106 Mid-Atlantic Ridge, Cenozoic sediment, 625 Mineralization, 94–6 Minnows, 568 Mississippi delta, copper, 593 Model for predictory copper toxicity, 567, 576 Molluscs, trace elements in shells used by paleoecologists, 574, 575 Mudskippers, 433, 435, 438, 439, 440, 441, 448, 452, 456, 459, 462 ” bitten by mosquitos, 462 Mullet, 544, 718 Mummichog, 545, 591, 719, 739 Muramic acid, indicator of bacterial biomass, 87 “Mussel Watch” programme used in detection of heavy metals in Western Port, Australia, 555, 556 Nancy Sound, N.Z., humic and fulvic acids, 608 Narragansett Bay, Rhode Island, copper, 556, 562, 588, 638 ” sediments, copper, 630 National Academy of Sciences, U.S.A., 175 Nektonic organisms and copper, 689–712 Neural control of luminescence in fishes, 498– 515 ” Myctophi-dae, 508– 10 ” Porichthys, 499–508 Neurotransmitter control, 503, 506, 514 New York Bight, sludge dumping, 539 ” City dumpsite, copper, 630 ” harbour, dredge-wastes, 626 ” Metropolitan area, copper and lead pollution, 560 ” pollution, 567 Nitrifying bacteria, 103, 104 Nitrilotriacetic acid (NTA), complexing of copper and zinc, 564 Noradrenaline, 505 North America, intertidal fishes, 420, 430, 431 ” Pacific coast, copper, 643 ” sandy-beach whelks, 358 ” Atlantic Central Water, 45, 57, 59 ” Deep Water, 59 ” pollution, 567 ” Pacific salmon, 32

Interlinking of physical

1428

” Sea, bacteria, 98 ” “classical” food web, 78, 79 ” copper, 588, 593, 594, 634, 638 ” intertidal fishes, 424 ” sources of copper, 585 ” West Pacific, coastal, copper, 643 Norway, copper, 597, 638 ” copper pollution and salmonid fisheries, 550 ” pollution in fjords, 554 Norwegian Sea, copper, 634 Nuclear testing fallout, 176 Nutrients, Antarctic, 8–19 OB, 1 Ocean quahog, 571 Ogee-chee River, copper, 629 Okhotsk Sea, copper, 629 Oligochaetes, abundance, 224–8 ” biomass and production, 224–34 Oligochaetes, Aeolosomatidae, 208 ” Aleutian Trench, 287 ” America, 242, 245, 266, 271, 297 ” α-amylase activity, 239 ” and alkalinity, 274 ” bacteria, 236, 237, 240, 241 ” grain size of substratum, 257 ” gregarines, 249, 252, 254 ” heavy metal pollution, 324– 5 ” light, 274–5 ” oil and oil dispersants, 326– 33 ” oxygen and hydrogen sulphide, 275–80 ” pollution, 321–33 ” redox potential discontinuity (RPD) layer, 275, 276, 277, 278, 279, 280, 282, 289, 304 ” salinity, 268–73 ” sediment agitation, 257– 9 ” sorting of sediment, 259– 60 ” temperature of substratum, 267–8 ” water content of substratum, 260–7 ” as hosts for parasites, 246–54 ” indicator species, 277 ” prey for carnivores, 241–5 ” Asia, 206 ” assimilation, 233–4 ” ATP-ase system, 312 ” Atlantic, 204, 206, 208, 225, 226, 232, 234, 245, 248, 299 ” average weights, 229, 230 ” Baja California, 265 ” Baltic, 208, 221, 222, 225, 226, 228, 229, 230, 231, 234, 237, 239, 242, 243, 244, 247, 253,

Oceanography and marine biology

955

260, 264, 266, 269, 271, 278, 282, 287, 288, 291, 292, 297, 300, 301, 302, 305, 307, 308, 309, 321 ” Bermuda, 267, 274, 299, 301, 302, 303, 310 ” Black Sea, 271, 291, 299, 326 ” biomass, 229–30; 231 ” biometrical data for ecological analysis, 202–3 ” bioturbation and trophic aspects, 281–3 ” California, 333 ” Canada, 282, 298 ” Cape Cod, 228, 232, 263, 287, 290 ” China, 208 Oligochaetes, culture methods, 203–4 ” Czechoslovakia, 297 ” Denmark, 254, 258, 276, 291, 312, 320 ” electrophoretic studies, 312 ” Enchytraeidae, 217–9 ” epipsammic browsing, 236 ” Europe, 208 ” extraction and evaluation of samples, 201–2 ” family diagnosis, 199–200 ” faunistic and zoogeographical aspects, 204–8 ” Finland, 269, 287 ” food, dissolved organic matter, 239–41 ” for Gobiidae, 242 ” pouting, 242 ” whiting, 242 ” fresh plant material, 239 ” live microalgae, 235 ” micro-organisms, 235– 8 ” of, 234–41 ” paniculate organic matter, 238–9 ” France, 224, 226, 228, 245, 247, 258, 269, 290 ” general occurrence, 286–7 ” Germany, 302, 305, 320 ” Great Barrier Reef, 206 ” Greenland, 208 ” habitat conditions and occurrence, 254–86 ” abiotic factors, 254– 80 ” biotic factors, 280– 6 ” substratum, 257–67 ” Iceland, 208 ” impact of domestic pollution, 321–3 ” industrial pollution, 323–33 ” India, 227, 228, 277, 297, 323 ” Indian Ocean, 204 ” Irish Sea, 226, 297, 305, 321 ” Italy, 247 ” Lake Esrom, Denmark, 233

Interlinking of physical

1430

” Lake Ontario, 233 ” Megascolecidae, 219 ” methods for biological and ecological studies, 200–4 ” Naididae, 208–12 ” North Sea, 225, 226, 231, 234, 239, 243, 244, 245, 252, 258, 259, 267, 268, 269, 271, 276, 279, 281, 283, 284, 289, 291, 295, 299, 302, 304, 305, 307, 309, 321, 326 ” Norway, 258, 290 Oligochaetes, occurrence, coralline and calcareous sands, 291 ” inland salt waters, 296 ” intertidal areas, 289–90 ” sandy beaches, 287–90 ” subtidal habitats, 290 ” oxygen consumption, 313, 314, 316, 317, 318 ” Pacific, 305, 309, 324 ” Pacific American coast, 206, 208 ” physiological mechanism and adaptation, 312–20 ” population structure, 219–24 ” preference experiments, 304– 12 ” production, 231–3 ” production and turnover, 234 ” reproduction, life cycle, and ” population structure, 208– 24 ” respiration rates, 314, 316, 317, 318 ” rôle in ecosystems, 333–7 ” Russia, 245, 247, 253, 302 ” salinity preferences, 305, 306, 307, 308, 309 ” salinity tolerance, 296–300 ” sampling, 200–1; 202 ” Scotland, 231, 238, 242, 245, 247, 252, 258 ” South Africa, 206, 261, 321 ” Soviet Pacific coast, 206 ” Spitsbergen, 247, 253 ” Sweden, 288, 302 ” temperature preferences, 305, 306, 308, 309 ” temperature tolerance, 300–3 ” Teseis oil spill, 326 ” tolerance and preference reactions, 296–312 ” trophic effects, 283–6 ” Tubificidae, 212–5 ” zonation patterns, 286–96; 335 Ontario, Canada, relation between metal levels and organic carbon in marshes, 578 Oregon State University, 134 Organic degradation, rôle of deep-sea fauna, 111–2 ” flux, estimation of from isotopic ratios, 84–5 ” matter, measurement of vertical flux, 82–4 ” mineralization, 94–6 ” produced by zooplankton, 78–81 Organic matter, supply in sea, 75–81

Oceanography and marine biology ” turnover in deep water by bacteria, 110–1 ” turnover in the sea, rôle of bacteria, 74–127 ” turnover rates by bacteria, 91–4 ” vertical flux in oceans, 81–5 ” particles, sinking rates, 81–4 Organo-metallic complexes and copper, 611 ” compounds, 739 ” and biological availability of copper, 604 oxaloacetate transaminase, 569 oxygenases, 498 Pacific coast, North America, intertidal fishes, 421 ” fur seal, 550, 613 ” intertidal fishes, 432, 443 ” Ocean, autochthonous organic matter, 83 ” bacteria, 86, 98, 104, 112 ” in surface film, 107 ” copper, 538, 613, 637, 638, 644 ” ferromanganese nodules, 624 ” flux of faecal pellets, 81 ” manganese nodules, 623 ” paniculate matter, 136, 142, 151 Pack-ice system, Antarctic, 26–7 Particle concentration, distribution of, 135– 6 ” measurement by continuous centrifuga-tion, 133 ” measurement by nephelometers, 134 ” measurement by optical methods, 134– 5 ” measurement by gravimetric methods, 130–4 ” measurement by transmissometers, 134– 5 ” measurement by water bottle samples, 130– 3 ” distribution spectra, 147–9 ” dynamics, 164–75 ” aggregation, 168–9 ” and radionuclide geochemistry, 175–90 ” dissolution, 169–71 ” first order scavenging processes, 171–4 ” fragmentation, 167–8 Particle dynamics, layering, 174 ” settling and deviation from Stokes’ Law, 164–6 ” size distribution, counters using light, 141 ” impedance-type counters, 141 ” in oceans, 136–55 ” sampling difficulties, 151–3 ” scanning electron microscopy, 142– 5 Particulate matter, estimation of fluxes, 153– 63; 155, 156, 157 ” large volume fil-tration systems, 159–63 ” sediment traps, 154–9 ” in oceans, measurement of particle concentration, 130–6

957

Interlinking of physical

1432

sampling methods, concen-tration, size distribution, and particle dynamics, 129–97 ” 210Po cycle, 185–9 ” settling velocities, 176, 177 ” two dimensional models for 210Pb, 185–9 ” metals, formation of, 616–8 ” organic carbon (POC), 78, 79, 84, 87, 99, 108, 115 ” estimation by pumping system, 82 ” estimation by sediment traps, 83 ” estimation by water sampling, 82 ” production by phytoplankton, 77 Patuxent River, copper, 597 Pee Dee River, copper, 629 Pelagic food web, hypothetical, 78, 79 Penpoint gunnel, 445 Peru-Chile upwelling system, 116 Peru upwelling, bacteria, 104, 106, 112, 113 Petrels, 570 Phentolamine, alpha-adrenergic agonist, 505 Philadelphia dumpsite, copper, 630 Photocyte, diagram of, 502 Photophore development, ultrastructural study, 507 ” structure and function in Argyropelecus, 522 Photophores, distribution of light from, 481, 483 Photosynthesis, effect of cadmium, 541 ” lead, 541 ” mercury, 541 Physical and biological processes in Antarctic Ocean, interlinking of, 1–35 Phytoplankton and copper, 644–5 ” as bioassay organism, 554 ” biomass and grazing by zooplankton, 78 ” primary production and excretion of organic matter, 76 ” production of dissolved organic carbon, 77 ” particulate organic carbon, 77 Pigeon guillemot, 461 Pinfish, 553, 718 Plaice, 446, 449, 547, 567, 713 Planktobacteria, 87 Planktonic organisms, copper in, 644–8 Polar front, 12, 29 ” ice cap, 2 Pony fish, 485 Portugal, intertidal fishes, 422 Port Valdez, Alaska, copper, 605 Predation of intertidal fishes, 461 Prokaryotic organisms, 88 Protozoa, effect of copper, 561 Prudhoe Bay crude oil, 588

Oceanography and marine biology ” effect on tidal flat, 605 Psychrophilic marine bacteria, 109 Puerto Rico, intertidal fishes, 422 ” oil-fired steam generating plant, 590 Puget Sound, intertidal fishes, 422, 430 ” rivers, copper, 630 Pulp-mill waste polluting sediments, 76 Pyruvate dehydrogenase, 566 Radionuclide geochemistry and inferred par ticle dynamics, 175–89 Radionuclides from testing of nuclear devices, 175–9 Rainbow trout, 546 Red abalone, 571 ” mangrove, accumulation of metals, 571 ” uptake of metals, 576 ” Sea blennies, 435 ” bream, 538 ” copper, 637 ” intertidal fishes, 443 ” metalliferous muds, 625 ” phytoplankton, 541 Red Sea tides, 609, 625 Reproduction of intertidal fishes, 436–41 Respiration of intertidal fishes, 435–6 Residence time, defined, 640 Resurrection Bay, Alaska, copper, 641 Rhode Island, U.S.A., pollution from electroplating plant, 576 Rhodotorulic acid, 739 Rhythmic behaviour, intertidal fishes, 453– 6 RNA, stained with Acridine Orange, 86 Ribulose 1, 5-bisphosphate carboxylase, 84 River Amazon, trace metals, 591, 592 ” Danube, copper, 629 ” Elbe, copper, 545, 597 ” Ems, copper, 641 ” heavy metals, 593 ” Liffey, Ireland, copper, 593 ” Rhine, copper, 592, 593, 594, 616, 629, 637, 641 ” Vistula, heavy metals, 593 ” Yukon, trace metals, 591, 592 Robert Koch’s philosophy, 74, 75 Rockall Trough, 131, Fig. 1 (Simpson) facing p. 131 Rockfish, 550, 612 Rockskipper, 452 Ross Sea, 26 Ross’ goose, 576 Rupert and Holberg inlets, B.C., copper, 641

959

Interlinking of physical

1434

Saanich inlet, B.C., cadmium and copper adsorption, 618 ” copper, 638, 641 St. Croix River, Maine, copper, 641 St. Johns River, copper, 629 St. Lawrence estuary, copper, 578 Salarine blennies in Red Sea, 443 Salinity, effect on intertidal fishes, 427–9 ” nutrient profile of Mediterranean outflow, 46, 47 Salinity-temperature-depth systems in Mediterranean outflow, 40 Salmon, 545, 567, 585 San Antonio Bay, Texas, dissolved metals, 626 San Diego Bay, copper, 597 ” Encina Power Plant, copper, 638 ” heavy metals, effect on kelp, 539 San Francisco Bay, copper, 594, 595, 616, 631, 641 ” dredging, 627 San Pedro Bay, copper, 641 ” dredging, 626 Sand goby, 425 Santa Barbara basin, copper, 631 Santee River, copper, 629 Sargasso Sea, copper, 635 ” copper and nickel, 538 ” sediment traps, 83 Satilla River, copper, 629 Savannah River, copper, 629 Scanning electron microscopy, bacteria, 87 SCOTIA, 1 Scotia Arc, 29 ” islands of, 5 ” Sea, 21 Sea of Okhotsk, intertidal fishes, 423 Sea urchin and effect of copper, 576 ” eggs in testing copper toxicity, 719 ” fertilization, effect of copper, 739 Sea-Tech transmissometer, 134 Seattle run-off into Puget Sound, copper, 631 Seaweeds as indicators of marine pollution, 648 Sediment, fossilized organic matter, 84 ” traps, 154; 83 ” design, 154 ” efficiency, 154–9 ” sample preservation, 159 Sei whale, 6, 7 Selenium, effect on Bullia, 405 SHACKLETON, 41 Shackleton cruise, 50

Oceanography and marine biology

961

” hydrographic data, 47, 51 Shanny, 442 Shark, 567 “Shore fishes”, 420, 421, 422 Sinking rates of organic particles, 81–2 Sockeye salmon, 538 South Africa, sandy beach, 351, 352, 355, 357, 358, 362, 366, 368, 369, 370, 371, 372, 374, 376, 388, 390, 394, 403, 407, 408, 409, 442 South African Museum, 408 ” America, intertidal fishes, 420 ” sandy-beach whelks, 352 ” Carolina, U.S.A., intertidal fishes, 422, 423 ” India, intertidal fishes, 422 ” Pacific, ferromanganese nodules, 623 Southern Ocean, 1, 9, 12, 19, 20, 21 ” ferromanganese nodules, 623 Soviet Academy of Sciences, 539 Starry flounder, 544 Starvation, effect on bacterial activity, 100–1 Strait of Gibraltar, Mediterranean water, 38, 39, 40, 41, 44, 45, 48, 53, 57, 58, 59, 60, 63, 67, 69 Striped bass, 546 “Subantarctic Mode” water, 21 Substrate-accelerated death, 101 Sulphate-reducing bacteria, 104, 105, 106 Superoxide dismutase, 537 Susquehanna River, copper, 537 Synergism between copper and nutrients, 738 Synergistic effect of copper and zinc, 544 Taiwan, intertidal fishes, 422, 423 Tamar estuary, U.K., copper, 637 TAMU nephelometer, 134 Taxonomy of marine bacteria, 85 Tees estuary, U.K., copper, 641 Temperature and intertidal fishes, 429–31 Temperature-salinity maxima in Mediterranean outflow, 50–8 profiles in Mediterranean outflow, 40, 42, 44, 45, 47, 48, 50, 51, 53, 57, 59, 65, 69 TERRA NOVA, 1 Territoriality of intertidal fishes, 452–3 Thames River, Long Is. Sound, copper, 630 “The Sanctuary”, whaling, 7 Thioplaca in sediments, 115 Thorium isotopes, scavenging models for, 180–5 “Tidal visitors”, 422 Timor Sea, fossil nodule, 624 Tokyo Bay, copper, 602 Toxicity of phenol to fish, 407 ” marine invertebrates, 408

Interlinking of physical

1436

” values, methods of evaluating, 716, 718 Triodothyronine, 429 Trophic rôle of marine bacteria, 98–100 Trout, 569, 573 Turbulent wave action, effect on intertidal fishes, 431–2 Tyrosinase, 537 UN EIFAC Working Group, 550 United Nations Joint Group of Experts on the Scientific Aspects of Marine Pollution (GESAMP), 623 U.S.A. municipal sludge, 587 U.S. Congress, 539 ” Environmental Protection Agency, 539, 742 ” Congress of Naval Research, 567 ” National Oceanic and Atmospheric Administration, 539 ” National Science Foundation, 1 ” Office of Naval research, 567 University of Cape Town, 352, 397, 404 ” Ecological Survey, 1952, 375 ” Port Elizabeth, 352 ” Washington, 134 Upper Newport Bay, Maryland, dumpsite, copper, 631 Upwelling regions, 78 Urophysical hormones, 429 Uruguay, intertidal fishes, 425 VALDIVA, 1 Vancouver, copper, 588 ” Is., copper mining, 625 Vertebrates, effect of copper, 566–9 Vertical eddy diffusion and upwelling velocity, 180 ” flux of organic matter in oceans, 81– 5 VITIAZ, 1 Wafra oil spillage, 406 WALTER HERWIG, 21 Water quality criteria, 742 ” copper, 741–3 Weddell Gyre, 29 ” Sea, 26, 28 West Indies, intertidal fishes, 422 ” Wind Drift, 9, 29 Western Port, Australia, heavy metals, 578 White Sea, copper, 629 Winter flounder, 544 Woods Hole Oceanographic Institution, 141 Woodward Biomedical Library, University of British Columbia, 536

Oceanography and marine biology

963

World War II, 29, 30 Yarra River, Australia, copper, 637 Yohimbine, alpha-adrenergic agonist, 505 Zebrafish, 532, 547 Zinc, effect on Bullia, 405 ” oxygen consumption of Bullia, 404 Zonation and habitat selection of intertidal fishes, 442–8 Zooplankton and copper, 645–8 ” biomass and phytoplankton production—a mismatch, 77, 78, 81 ” production of organic matter, 78–9