Advances in Ecological Research, Volume 21

Advances in

ECOLOGICAL RESEARCH VOLUME 21

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Advances in

ECOLOGICAL RESEARCH Edited by

M. BEGON Depnrtmenf of Zoology, University of Liverpool, Liverpool, L69 S B X , U K

A. H. FITTER Department of Biology, University of York, York, YO1 5 D D , U K

A. MACFADYEN 23 Mountsandel Road, Coleraine, Northern Ireland

VOLUME 21

ACADEMIC PRESS Harcourt Brace Jovanovich, Publishers London San Diego New York Boston Sydney T o k y o Toronto

ACADEMIC PRESS LTD. 24/28 Oval Road London NWl 7DX United Stares Edition published by ACADEMIC PRESS INC. San Diego, CA 92101

Copyright @ 1991 by ACADEMIC PRESS LIMITED

All Rights Reserved No part of this book may be reproduced in any form by photostat, microfilm, or any other means, without written permission from the publishers

British Library Cataloguing in Publication Data Advances in ecological research. VOl. 21 1. Ecology 1. Begon, Michael 574.5 ISBN 0-12-013921-9

This book is printed on acid-free paper Editorial and production services by Fisher Duncan 10 Barley Mow Passage, Chiswick. London W4 4PH, UK Printed in Great Britain by St Edmundsbury Press Limited, Bury St Edmunds, Suffolk

Contributors to Volume 21 L. I . BABBAR, The University of Michigan, Ann Arbor, M I 48109, USA. M. BRUNDREIT, Soil Science and Plant Nutrition, The University of Western Australia, Nedlands, W A 6009, Australia. C . P. CHANWAY, Department of Forest Sciences, University of British Columbia, Vancouver, British Columbia, Canada V6T 1W5. A . J . GRAY, Institute of Terrestrial Research, Furzebrook Research Station, Wareham, Dorset BH2 5AJ, U K . F . B. HOLL, Department of Plant Science, University of British Columbia, Vancouver, British Columbia, Canada V6T 2 A l . D . F . MARSHALL, Institute of Terrestrial Research, Furzebrook Research Station, Wareham, Dorset BH2 5AJ, UK. A. F . RAYBOULD, Institute of Terrestrial Research, Furzebrook Research Station, Wareham, Dorset BH2 5AJ, UK. D. D . RICHTER, School of Forestry & Environmental Studies, Duke University, 214 Biological Sciences Building, Durham, NC 27706, USA. R. H. SMITH, Department of Pure and Applied Zoology, University of Reading, Whiteknights, PO Box 228, Reading RG6 2AJ, UK. R. TURKINGTON, Department of Botany, University of British Columbia, Vancouver, British Columbia, Canada V6T 2 B l .

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Preface This volume of Advances in Ecological Research contains five papers, two of which have an evolutionary slant, and two of which are concerned with plant-microbe interactions. The fifth paper, by Richter and Babbar reviews the diversity of tropical soils. There has long been a failure of communication between soil scientists and ecologists, made more serious by the simultaneous use of four different soil science taxonomies with numerous subsidiary dialects. The consequences have been particularly serious in the field of tropical ecosystem studies, where concepts developed outside the tropics have sometimes forcibly been imposed on unfamiliar soil types. One consequence of Richter and Babbar’s approach to clarifying the situation will certainly be that adherents among the soil science fraternity who espouse one of the existing systems will object to “oversimplification” and other misinterpretations of their discipline. If this article reveals that the emperor has n o clothes and stimulates discussion, it will achieve a useful function and may lead to a clearer view of the diversity of tropical soils. Certainly there is a need for soil scientists and ecologists to develop a better working relationship, as it is increasingly recognized how many fundamental ecological processes occur in the soil. Both Brundrett and Chanway, and his colleagues, examine aspects of the interactions between plants and soil micro-organisms. The mycorrhizal symbiosis is certainly the most widespread of all symbioses, but although a good physiological explanation of the function of most types of mycorrhiza has been achieved, their ecological significance is much less clear. Brundrett’s wide-ranging review pulls together a very diverse literature and points to important patterns in the mycorrhizal relationship on a global scale. Interactions between plants and other soil microbes are even less well understood, the Rhizobiurn symbiosis excepted. Chanway, Turkington and Holl look at a number of these interactions and consider the significance of specificity in determining their ecological behaviour. There are a priori expectations as to the relationship between the extent of mutualism and of specificity in these interactions, but the ecological consequences are less clear. There are a number of intriguing results in the literature, which are discussed in this article, which suggest that a vii

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degree of specificity may occur in some of these interactions and may have important effects on plant-plant competition. Chanway and colleagues point to an important future area of research activity. The application of evolutionary ideas to such areas as this seems likely to bring considerable benefits; in evolutionary ecology in general, it is perhaps the study of life history patterns that has seen the greatest intellectual activity over the last twenty years. Nevertheless, a number of areas of confusion and deep uncertainty remain, including the questions of whether attention should be focussed on genotypes or phenotypes, and how these are likely to interrelate. In his paper, Smith makes great steps towards dispelling this confusion, by a careful and balanced exposition of the underlying ideas, and by a close look at one particular experimental system, from which he and his colleagues have obtained extensive and very pertinent data. One of the central questions of evolutionary biology is in practice an ecological one: how do species originate? One of the few documented cases of speciation in the last hundred years is that of Spartina anglica, the saltmarsh grass which has been so spectacularly successful at colonizing previously unoccupied regions of salt marshes in many temperate regions of the world. In effect, it appears to have occupied a vacant niche. Fortunately, its origin and spread have been carefully documented, and Gray, Marshall and Raybould give the first detailed account of its origin, spread and ecology. One striking feature that emerges is its extraordinary lack of genetic variability, which seems to have important implications for its future status and ability to withstand pathogens, as well as allowing its niche to be defined with remarkable precision. These papers form a group that represent several actively developing areas of ecology, but have no common theme. We make no particular effort to ensure that individual volumes of Advances in Ecological Research have such a theme, though we may from time to time put together special volumes on especially topical themes. One such that will shortly be appearing is on Global Change and is edited by a guest editor, Professor Ian Woodward. If readers have ideas on topics that would be appropriate to such treatment, we will always be happy to consider them. A . H. Fitter M. Begon A . Macfadyen

Contents Contributors to Volume 21 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Preface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

v

vii

A Century of Evolution in Spartina anglica A . J . GRAY. D . F . MARSHALL and A . F . RAYBOULD

I . Summary .......................................................... I1. Introduction 111. The History of Spartina anglica .................................... A . The Discovery and Spread of Spartina x townsendii B . The Discovery and Spread of Spartina anglica .............. IV . The Origin of Spartina anglica A . Early Ideas on the Origin B . Evidence for the Origin C . "Spartinu x neyrautii" D . Problems over the Origin V . Variation in Sparrina anglica ...................................... A . Sources of Genetic Variation in Spartina anglica B . Evidence for Genetic Variation in Spartina anglica ......... VI . The Ecology of Spartina anglica A . Rates and Pattern of Spread and Dispersal B . Growth and Production C . The Niche of Spartina anglica D . Interactions with Other Species E . Die.back. Control and Conservation VII . The Future of Spartina anglica ..................................... Acknowledgements ...................................................... References

....................................................... ........ ..................................... .................................. .................................... ..................................... .................................. ........... ................................... ................. .................................... .............................. ............................ .......................

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1 3 4 4 7 10 10 12 20 21 21 21 23 27 27 30 35 39 45 49 51 51

Genetic and Phenotypic Aspects of Life-history Evolution in Animals

R . H . SMITH I . Summary .......................................................... I1 . Introduction ....................................................... A . The Pieris rapae Problem .................................. B . Aims of This Review

...................................... ix

63 64 65 67

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C'ON7'ENTS

111. Conccptual Framcworks ........................................... A . Lifc-history Theory ........................................ B . Quantitative Genetics., .................................... C . Multivariatc Selcction ...................................... D . Gcnc-Environment Interactions ............................ E . Thc G and E Matrices ..................................... 1V . Experimental Approaches Illustrated by CNllosohrirc/iiis ........... A . Phenotypic Correlation .................................... B . Experimental Manipulations ............................... C . Breeding Designs .......................................... D . Selection Experimcnts E . Mapping the Options Set .................................. V . Discussion ......................................................... Acknowledgements., .................................................... References ..............................................................

.....................................

68 69

77 83 88 91

93 94 97 100 106 109 110

112 113

Ecological Implications of Specificity Between Plants and Rhizosphere Micro-organisms

C . P . CHANWAY. R . 'TURKINGTON. and F . B . HOLL I . Introduction. ...................................................... 122 I1 . Specificity Betwcen Plants and Beneficial Micro-organisms ......... 127

111.

R/l;:oh;~ittz

........................................................

.................................................. .............

A . Infectivity B . Effectiveness ............................................... C . The Relationship of Infectivity to Effectiveness

129

129 133 135

........................................................... 136 A . Infectivity .................................................. 137 B . Effectiveness ............................................... 138 V . Associative Rhizosphere Bacteria .................................. 139 139 A . Infectivity .................................................. B . Effectiveness ............................................... 140 143 VI . Ectomycorrhizal Fungi ............................................ A . Infectivity .................................................. 144 B . Effectiveness ............................................... 144 VII . Vesicular-Arbuscular Mycorrhizal (VAM) Fungi .................. 146 A . Infectivity .................................................. 146 B . Effectivencss ............................................... 147 VI11 . Effects on Plant C'ompctition ...................................... 147 A . Symbiotic Bacteria ......................................... 147 150 B . Associative Rhizosphere Bacteria .......................... C . Mycorrhizal Fungi ......................................... 152 IX . Consequences of Specificity on Plant Community Structure ........ 153 X . Conclusions ....................................................... 156 Acknowledgements ...................................................... 157 liefercncos .............................................................. 157 IV .

Frritikiri

CONTENTS

Xi

Mycorrhizes in Natural Ecosystems

.

M BRUNDRETT 1. Summary .......................................................... I1 . Introduction ....................................................... 111. Mycorrhizal Ecology .............................................. A . Mycorrhizal Fungi ......................................... B . Edaphic or Climatic Factors and Mycorrhizal Fungi ........ C . The Host Plant ............................................ D . Plants and Mycorrhizal Fungi .............................. E . Edaphic/Environmental Factors. Plants and Mycorrhizas ... F . The Ecology of Mycorrhizal Plants ......................... IV . Conclusions V . Appendix 1 ....................................................... Acknowledgements ...................................................... References ..............................................................

.......................................................

171 172 173 175 176 194 196 202 235 257 262 271 271

Soil Diversity in the Tropics

D . D . RICHTER and L . I . BABBAR I . Introduction and Objectives ....................................... I1 . Changing Perspectives about Soil Taxonomy ....................... 111. The Development of Misconceptions about “Tropical Soil” ........ A . The Enormous Challenge of Mapping Soils on the 5-Billion Hectare Tropical Landscape ............................... B . Interdisciplinary Miscommunication about Soils in the Tropics .................................................... C . The Tower of Babel Effect of Too Many Taxonomies and Nomenclatures ........................................ D . Emphasis on Factors Rather than on Effects of Soil Formation: The 1938 Soil Classification .................... IV . Advances in Soil Taxonomy and the Creation of the World Soil Map ......................................................... A . Soil Taxonomy: A New Scientific Paradigm ................ B . FAO/UNESCO Soil Map of the World V . Diversity of Soil Taxa in the Tropics .............................. A . General Soil Taxonomic Variation in Tropical Africa, America and Asia ......................................... B . Ferralsols: Modern Inheritors of the Latisol Concept ...... C . Acrisols: the Underestimated Soil Order .................. D . Lithosols. Arenosols and Luvisols: From Extremely Fertile to Infertile. 500 Million Hectares Each .................... E . Regosols. Yermosols and Cambisols: Weakly Developed Soils ...................................................... F . The Other 1 Billion Hectares: Extreme Variation ..........

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316 320 322 332 333 333 336 342 342 346 350 356 357 359 360 361 362

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CONTENTS

VI . How Much Area in the Tropics Is Covered by Oxisols?: Results of the First Soil Surveys of the Amazon Basin A . Background ............................................... B . Some Common Properties of Amazonian Soils ............ C . The New Soils Map of the Brazilian Amazon .............. D . How Much Area in the Tropics Is Covered by Oxisols? .... E . A Realistic Concept of Soil Diversity in the Humid Tropics VII . Meso- and Local-scale Soil Variation in the Tropics VIII . Conclusions ...................................................... Acknowledgements ...................................................... References

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...................................................

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Index ................................................................... Cumulative List of Titles ................................................

367 367 369 370 377 378 370 381 382 383 391 40 I

A Century of Evolution in Spartina anglica A . J . GRAY. D . F . MARSHALL and A . F . RAYBOULD

I . Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111. The History of Spartina anglica ........................ A . The Discovery and Spread of Spartina X townsendii . . . . . B . The Discovery and Spread of Spartina anglica IV . The Origin of Spartina anglica ........................ A . Early Ideas on the Origin ....................... B . Evidence for the Origin ........................ C . “Spartina x neyrautii” ........................ D . Problems over the Origin ....................... V . Variation in Spartina anglica ......................... A . Sources of Genetic Variation in Spartina anglica . . . . . . . B . Evidence for Genetic Variation in Spartina anglica VI . The Ecology of Spartina anglica ....................... A . Rates and Pattern of Spread and Dispersal . . . . . . . . . . . B . Growth and Production ........................ C . The Niche of Spartina anglica .................... D . Interactions with Other Species . . . . . . . . . . . . . . . . . . . E . Die.back. Control and Conservation . . . . . . . . . . . . . . . VII . The Future of Spartina anglica Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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I SUMMARY This review describes the origin and spread of the salt marsh grass Spartina anglica. the first specimen of which was collected from Lymington. Hampshire in 1892. and presents new material relating to its ecology and evolution . Using electrophoretically detectable variation. we confirm the strong circumstantial evidence that S . anglica originated by Copyright 0 1991 Academic Press Liniired All rights of reprodrrcrion i n any form reserved

ADVANCES IN ECOLOGICAL RESEARCH VOL . 21 ISBN 0-12-013921-9

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A . J . GRAY E T A I . .

chromosome doubling of the sterile hybrid between the Old World species Spartina maritima and the North American Spartina alterniflora . In addition, extensive analysis of isoenzyme and seed protein phenotypes indicates that S . anglica is almost totally lacking in genetic variation. This may result from a narrow genetic base following a single origin or from a multiple origin from the uniform parents. It is likely to be maintained by the extensive clonal spread of most populations from a few, often deliberately introduced, founders and by preferential pairing at meiosis between identical homologous chromosomes preventing recombination between the component genomes. The implications of these findings are manifold. For example, the genetic uniformity of the species may help to explain why it has a relatively narrow ecological amplitude. A simple multiple-regression model incorporating largely physical, tide-related variables indicates that the distributional limits of Spartina in metres above Ordnance Datum are predicted remarkably well by tidal range, with variation in fetch, estuary area and position on the estuarine gradient significantly improving the prediction (generating equations explaining more than 90% of the variation in both the upper and lower limits). That the niche of the species can be so well defined is most probably due to its recent evolution and its lack of genetic differentiation as well as the predominance of physical, as opposed to biological, factors limiting its downshore spread. The species’ genetic uniformity, coupled with its frequent occurrence as dense, monospecific stands, may also account for the recent rapid spread in several populations of the ergot fungus, Claviceps purpurea. For example, in Poole Harbour, Dorset, the average level of infected inflorescences rose from 36% in 1983 to more than 85% in 1988. Spartina anglica is, unusually among temperate grasses, a C4 species (one of only eight such species in Britain), and the implications of this method of carbon fixation in a species whose biomass production and possible northward spread may be limited by early spring and summer temperatures is considered, particularly in relation to projected climatic warming. The short-term causes and consequences of die-back in the southern parts of the plant’s range are described, including the changes in low water tidal channels accompanying the invasion and subsequent decline of Spartina in a south coast harbour. In a section describing the species’ interaction with other species we discuss the effects of Spartina’s invasion on aerial and benthic invertebrates and on overwintering wading birds, particularly the dunlin, Calidris alpina, the decline in numbers of which correlates with the spread of the grass in British estuaries. Soil conditions, grazing and temperature are important factors affecting the competitive interaction

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of S. anglica with Puccinellia maritima, the latter invading S. anglica swards more rapidly in sandier, grazed marshes in more northern latitudes. The consequences of the genetic bottleneck that occurred during the speciation process and the relative youth of S. anglica in evolutionary terms are emphasized throughout by comparing the species with P. maritima. The greater variation, population differentiation and niche breadth of the latter species are particularly evident. Finally, we look forward to the changes which may occur with global warming and rising sea levels and how they may affect the second century of the species' evolution.

11. INTRODUCTION The perennial salt marsh grass Spartina anglica C.E. Hubbard has become the textbook example of allopolyploid speciation and successful invasion by a new species (Fig. 1). First recorded from Lymington, Hampshire, in 1892, it now occurs in temperate zones throughout the world, mainly because it has been widely planted to stabilize tidal mud flats. It has profoundly altered the ecology of mud flats and salt marshes over many thousands of hectares.

Fig. 1. S. anglicn.

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A . J . GRAY ET A L .

The first one hundred years of the “Spartina story” (Lambert, 1964) have produced an extensive literature -a bibliography with more than 480 references was compiled by J.C.E. Hubbard more than 25 years ago (Hubbard, 1965, unpublished). In this contribution we do not attempt a comprehensive review of that literature, but highlight the important landmarks in the study of the species. These include its discovery and spread in the early part of the century, the detective work in the late 1950s and early 1960s on its origins, the extensive research during the same period on its ecology, particularly in southern England where S. anglica-dominated marshes were already showing extensive die-back, and the demonstration in 1975 that the species utilizes the C4 photosynthetic pathway. These findings are discussed in the light of recent significant advances in several research areas. Among the most illuminating of these are biochemical confirmation of the species’ origins and the evidence of a narrow genetic base. The latter is particularly relevant in considering the development of the species’ niche, its interaction with other species, especially pathogens, and its future evolution. The following sections bring together these recent findings and attempt to demonstrate that S. anglica continues to provide a rare and fascinating opportunity to gain insight into the early stages of species evolution. In examining this continuing and dynamic process we also look forward to the possibility of cycles of invasion and decline, of northward spread under global warming, and of changing relationships with pathogens and competitors.

111. THE HISTORY OF SPARTZNA ANGLZCA A. The Discovery and Spread of Spartina

X

townsendii”

In 1879, H. and J. Groves described a plant growing on mud flats at Hythe, Hampshire “. . . which, by the majority of characters given in British works, would be S. alterniflora rather than S. stricta [= S. rnaritirna], yet which we now consider to be the latter”. By 1880, however, the Groves had revised their opinion that the plant was a form *Nore on nomenclarure: For many years it was thought that there was only one form of the hybrid between S. marifima and S. alrerniflora. namely the fertile amphidiploid. In 1957, however, C.E. Hubbard showed that there still existed the F1 hybrid between S. maririma and S. alterniflora. H e stated that this form was the one described by the Groves, and. therefore, the name “rownsendii” should only be applied t o the initial hybrid rather than its amphidiploid derivative. This left the amphidiploid without a name. Hubbard (1968) invalidly published the name S. anglica for the amphidiploid. although later (in Heywood. 1978, p. 364) he published a valid diagnosis. so legitimizing the name. The term ”S. ariglica agg.” will be used in this article when referring specifically to both forms together.

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of Spartina stricta and wrote that it “. . . occupies so intermediate a position between that species and S. alterniflora, that it appears desirable not to include it in either”. The Groves (1880, 1882) named this plant Spartina fownsendi (spelt initially with one “i”; the second was added by other authors and is now the accepted spelling) in honour of Frederick Townsend, author of the Flora of Hampshire (Britten, 1906). There is some dispute as to how long these forms of Spartina had been at Hythe before the Groves described them as a new species. According to Hubbard (1965a), the Groves collected the plants on which they based their descriptions in 1877. Hubbard states, however, that a Mr R.S. Hill had collected specimens of the same plant at Hythe in 1870, but that the plant was thought to be a luxuriant form of Spartina rnaritima. Sutherland and Eastwood (1916) also gave 1870 as the first date of collection of S . X townsendii based on a specimen in the Warner Herbarium, University of Southampton. They suggest, moreover, that descriptions of Spartina alterniflora from near Southampton given by Sowerby in his “Grasses of Great Britain”, published in 1861, are very close to those of S. x townsendii published by the Groves. The earliest confirmed date for the discovery of S. x townsendii, however, remains 1870. The plant spread relatively slowly at first and had not reached Cracknore Hard (2 miles north of Hythe) by 1883 (Stapf, 1908), but was recorded from Southampton in 1887 (Stapf, 1913). Stapf (1913), however, describes how “. . . towards the end of the eighties, something occurred that favoured the spreading of the grass”. Marchant (1967) correlates this with the formation of the seed-bearing amphidiploid S. anglica. Indeed, the first confirmed record for S . anglica is from 1892 at Lymington, Hampshire. After this date it often becomes difficult to know whether records are referring to S . X townsendii or S . anglica as the distinction between the two was not recognized until 1957 (see footnote p.4). Goodman et al. (1969) state that sterile material features largely, though not exclusively, in collections from 1892 to 1910, such as those from Yarmouth, Isle of Wight (1893), Poole Harbour (1905), and Lymington (1893 and 1907). After 1910 it is likely that as S . anglica was deliberately introduced into other parts of the British Isles for coastal protection and land reclamation (see below). Spartina X townsendii was also introduced into some areas due to mixed collections of both species being made, especially as many collections were made from Poole Harbour where S. x townsendii was still present in several sites in the 1960s (Hubbard, 1965a). Spartina x townsendii is still present as substantial swards at Hythe and has been found on the landward edge of S. anglica swards in Norfolk (Swann, 1965), Poole Harbour (Hubbard, 1965a) and the

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A . J . GRAY E T A L .

Bristol Channel (Holland, 1982). The preference of S . x rownsendii for these higher, more stable habitats also occurs in Holland (Drok, 1979). Hubbard and Stebbings (1967) estimated that the total area covered by S. x townsendii in Britain was only 20 ha compared with about 12000 ha of S. anglica. As in the British Isles, S . anglica has been used overseas for land reclamation and coastal defence, and it seems inevitable that some S. x townsendii has been introduced along with it, especially when the material had been taken as cuttings from Poole Harbour. Spartinu x townsendii has been recorded from Holland on the basis of cytological data (Drok, 1979, 1983) and may also occur in New Zealand. The presence of S. x townsendii there is based on the fact that much " S . anglica" in New Zealand produces no seed (Allan, 1930; observations of C. E. Hubbard in Ranwell, 1967); however, the erratic nature of seed set in S . anglica (see below) makes this a poor diagnostic character. The other confirmed site for S . x townsendii is in the extreme south west of France and north east of Spain around Hendaye and San Sebastian, although it has only just been recognized as such. In 1892 Neyraut collected an unusual form of Spartina at Hendaye and these plants were named S . x neyrautii by Foucaud in 1894 (Mobberley, 1956). This species has long been regarded as a hybrid between S. maritima and S . afterniflora and recent evidence from isozyme studies has confirmed this (see below). Under the International Code of Botanical Nomenclature, therefore, S. x neyrautii must be regarded as a synonym of S. x rownsendii as it was the later of the two names to be published (Raybould et al., 1990). Although S . x townsendii occurred in several locations in this area, extensive land reclamation reduced it to a single site near San Sebastian Airport (Hubbard et al., 1978). Another complication when considering the distribution of S . x townsendii is the occurrence of polyhaploids derived from S. anglica. Marchant (1967) described a seedling of S . anglica that was morphologically very close to S . X townsendii rather than its S . anglica parent. A chromosome count of this plant gave 2n = 61, which was exactly half the number of the parent clone (Marchant, 1968). This seedling was produced presumably by the development of an unfertilized egg cell (agamospermy), and its similarity to S . x townsendii is to be expected as homologous chromosomes in S. anglica are identical and chromosome pairing is predominantly as bivalents (see below). Spartinu anglica gametes, therefore, will be very similar genetically to S . x townsendii. Marchant (1975) has suggested that S . x townsendii in some parts of the country may be secondarily derived from S. anglica by polyhaploidy. The occurrence of dwarf, densely tillering forms of S . anglica that have been described from the Bristol Channel and the Dovey Estuary (Chater

A CENTURY OF EVOLUTION IN SPARTINA ANGLICA

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and Jones, 1951; Chater, 1965), from areas north of Dublin (Boyle, 1976a), south west Holland (Drok, 1979, 1983) and New Zealand (Allan, 1930; Bascand, 1970) may represent polyhaploid clones; careful cytological examinations are required, however, to confirm this suggestion.

B. The Discovery and Spread of Spartina anglica As stated above, the first record for S . anglica is from Lymington in 1892 and after this date there is uncertainty as to whether records refer to the sterile F1 or the amphidiploid. The exact place and date of the origin is unknown. Although the rate of spread in the 1890s and 1900s was rapid, suggesting spread by seed, Goodman et al., (1969) believe that in many areas the F l was predominant. Whatever the proportions, by 1899 either S. anglica or S . X townsendii had begun to colonize Poole Harbour in the west and Chichester Harbour in the east, and many intervening areas (Stapf, 1908). On the north coast of the Isle of Wight there were populations at Yarmouth by 1893, on the River Medina in 1895, and on several marshes between Ryde and Cowes in 1907 (Stapf, 1908). Back on the mainland, Christchurch Harbour was beginning to be colonized by 1913 (Oliver, 1920) and Pagham Harbour by 1918 (Oliver, 1925). Oliver (1925) also gives a record from the mouth of the River Rother at Rye. The colonization of salt marshes between Poole and Rye probably represents the extent of completely natural spread of S . anglica in the British Isles, although extensive secondary natural spread following deliberate introductions has occurred (e.g. Goodman et al., 1959). Spartina anglica also spread to the north coast of France by natural means (or by accidental spread through shipping activity; there is no record of S . anglica ever having been deliberately introduced into France (Ranwell, 1967)). The first records are from Reville, on the north east tip of the Cotentin Peninsula, and Carentan at the mouth of the River Vire in 1906; slightly further east, the species was found at Sallenelles on the River Orne just north of Caen in 1918 (Oliver, 1925). By 1926 there were 1000 ha of S . anglica at Carentan (Oliver, 1926). The other area where S. anglica has been very successful in France is on the marshes at the mouth of the River Seine. The first record is from Tancarville in 1915 and from Harve and Deauville in 1922 (Oliver, 1925). Oliver (1926) says that the rate of spread of S . anglica on the Seine marshes was the fastest he had ever seen, and Senay (1934) estimated that the plant covered 1000 ha. Spartina anglica seems to have colonized many of the estuaries on the northern coast of France and

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A . J . GRAY E T A L .

Ranwell (1967) estimated that there were between 4000 and 8000 ha of Spartina by the mid-1960s. The distribution of S. anglica, both in the British Isles and around the world, has been radically altered by deliberate introductions of the species. In the 1900s it was realised that S. anglica was very efficient at stabilizing bare mud flats (Stapf, 1907). This led to its utilization by private landowners as a method of reclaiming mud flats for agriculture and by coastal defence authorities for coastal protection. Many introductions of the species have been made around Great Britain and they have been thoroughly catalogued by Goodman et al. (1959) and Hubbard and Stebbings (1967). The first record of deliberate planting of S. anglica is from the Beaulieu Estuary in 1898, and this was followed by north Norfolk in 1907 and the Lincolnshire Wash in 1910. Most of the planting, however, was carried out in the 1920s (e.g. Bryce, 1936) with major efforts (utilizing 2000 or more cuttings (Ranwell, 1967)) in the Thames and Blackwater Estuaries, the Exe, the Severn Estuary (especially on the Somerset coast), the Dee, and the Humber. The east coast plantings were very successful, and during the 1930s virtually every estuary in Essex and southern Suffolk was either planted with Spartinu or was colonized as the result of natural spread from the introduced material (Goodman et al., 1959). The other major population of S. anglica on the east coast is in the Wash with further smaller populations as far north as Udale Bay in the Cromarty Firth (Smith, 1982). The plant has also been introduced successfully on the west coast. Hubbard and Stebbings (1967) record that the estuaries of the Severn, the Burry, the Dovey, the Mawddach, the Dee, the Ribble, and the Wyre all have populations of S. anglica in excess of 100 ha. Numerous other estuaries and salt marshes between these major sites have S. anglica populations, with the furthest north being the Isle of Harris in the Outer Hebrides. Goodman et al. (1959), Hubbard and Stebbings (1967) and Ranwell (1967) all estimated the total area of S. anglica in Great Britain to be about 12000 ha. This has been revised recently by Charman (1990) to 10000 ha, though Charman points out that part of the disparity may lie in differing estimation procedures. What is clear, nevertheless, is that there have been significant changes in the distribution of S. anglica since 1967. Charman records a 44% reduction in cover on the east coast, an 11% reduction on the south coast, and a 40% increase on the west coast. The south and east coast populations of S. anglica are some of the oldest, whilst those on the west coast are much younger. Indeed, Charman reports that 20 new S. anglica populations have become established on the west coast since Hubbard and Stebbings’ survey. This is evidence that S. anglica populations undergo a change from invasion and sward formation to regression (often termed “die-back”) . Reasons

A CENTURY OF EVOLUTION IN SPARTINA ANGLICA

9

for this are discussed in later sections. The current distribution of S. angfica in Great Britain is shown in Fig. 2. In Ireland, S. angfica was first introduced on the south coast in the River Lee in 1925 (Cummins, 1930). Further introductions were made on the west coast in the Shannon and Fergus in 1928, 1931 and 1932

Fig. 2. The distribution of S. anglica (including records of Spartinu X townsendii) in the British Isles (GB-204, Ir-27, Ch.Is-1). Map supplied by the Biological Records Centre, ITE, Monks Wood.

10

A . J . GRAY E T A L .

(Praeger, 1932). On the east coast, introductions were made in the Dublin Bay area in 1932 (Boyle, 1976a) and in Belfast Lough in 1929 (Praeger, 1932). Ranwell (1967) estimated that there were between 200 and 400 ha of S. anglica in Ireland. Spartinu angfica has now been successfully introduced into many areas outside the British Isles, notably Holland (4000-5000 ha), Germany (400-800 ha) and Denmark (500 ha), with smaller introductions in Australia and New Zealand (Ranwell, 1967). The most remarkable spread has been that following the introduction of the species to China. The descendents of only 21 individuals, survivors of a batch of 35 sent as seed in thermos flasks by D. S. Ranwell in 1963, have spread, mainly by planting, along almost the entire Chinese coast to occupy, by 1980, more than 36000 ha (Chung, 1990). Spartina anglica's present geographical range is from 57.5 "N to 48 "N in Europe, 46 "S to 35 "S in New Zealand and Australia (Ranwell, 1967), and from 41 ON tQ 21 ON in China (Chung, 1990). Unsuccessful introductions of S. angfica have been made on the eastern seaboard of the USA, in the Caribbean and the Guianas, India and South Africa. Ranwell (1967) has suggested that winter temperature may be the crucial factor in whether establishment is successful; the species is damaged by frost, and warm winter temperatures are said to impede development. It is worth remembering, however, that S. anglica survived for nearly 2 years in British Guiana and was only exterminated after it was overrun by a native Spartinu species, Spartina brasifiensis (Lambert, 1964). To conclude, it is interesting to note the role of Poole Harbour in the spread of S. angfica, especially in the light of the scarcity of genetic variation in the species (see below). Hubbard (1965a) estimated that between 1924 and 1936 over 175000 cuttings were exported from material derived from Poole Harbour, and of these 46000 were planted in the British Isles. This figure does not include numerous batches of seed from this material that were also dispatched. This must represent a severe genetic bottleneck for the species.

IV. THE ORIGIN OF SPARTZNA ANGLZCA A. Early Ideas on the Origin Despite the fact that the Groves' descriptions of S. X townsendii had emphasized that the plant had many morphological characters intermediate between those of S.maritima and S. alternifloru, it is surprising how little the possibility of a hybrid origin is considered in the early literature. Most authors tried to fit the new form into previously

A CENTURY OF EVOLUTION IN SPARTINA ANGLICA

11

decribed taxa. Boswell (in Groves and Groves 1882) suggested that the plant might be ”true” Spartina glabra, and Townsend (1883) thought that the plant was a variety of S. maritima. Corbibe and Chevalier, writing in 1906 about the appearance of Spartina in Normandy, claimed that S. angfica could not be a hybrid as it had fertile pollen and seeds, and because S. alterniflora did not occur in northern France (Oliver, 1925); they thought the plant to be S.glabra var. pilosa. Many of the early explanations of the appearance of the S . angfica agg. were based on the possibility of a direct introduction of foreign Spartina taxa, usually one of the numerous named varieties in the S . alterniflora/S. glabra complex. This is probably more a reflection of the confused state of the taxonomy of S. alternifloru at that time than to any great resemblance of the S . anglica agg. to the published descriptions of other taxa (see Mobberley (1956) for a review of the ideas on the taxonomy of S. alterniflora). The introduction of a foreign species theory was disposed of by Stapf (1908), who found that no species of Spartinu exactly matched the forms that had first appeared at Hythe. Stapf proposed that the most plausible alternative was that the S. anglica agg. had arisen by hybridization between S . maritima and S. alterniflora, given the intermediate morphology of the plants and their similarity to “S. x neyrautii” which had arisen in the only other location where S. maritima and S. alterniflora co-existed (Stapf, 1908). This idea rapidly became accepted, although the uniformity of seed progenies of S . anglica created problems because if the species was a hybrid, segregation of parental traits would be expected. The situation was summed up by Oliver (1925) who wrote: If it [ S . anglica] is a hybrid, then it should betray its hybrid constitution. In one respect it does this, viz. in the extreme vegetative vigour which it displays. A vigorous constitution of this kind is not unusual in a first cross; what is surprising is that not only is this vigour maintained through many seed generations but that the plant should remain substantially uniform. There is no evidence of segregation, which is to be expected in an ordinary hybrid.

Oliver goes on to say that the lack of segregation may not necessarily be a problem and appears to propose the possibility that the characteristics of s. anglica could be explained by allopolyploidy: This circumstance [lack of segregation], which at first sight seems to negative the hybrid theory, does not however absolutely close the door to i t , because there exists a class of hybrids characterized by perfect stability. But it would be premature to assert that S . townsendii belonged to this class, without the production of evidence.

Huskins (1930a,b) finally provided this evidence when he made chromosome counts on S. anglica and its putative parents, and found them to

12

A . J . G R A Y ET A L .

be in perfect agreement with the theory of an allopolyploid origin. Although Huskins’ figures have proved to be inaccurate (see below), the publication of his papers firmly established allopolyploidy as the prevailing theory to account for the origin of S . anglica, and this theory has never been in serious doubt. There is now a wealth of evidence to show that S . x townsendii is a hybrid between S . maritima and S . alterniflora, and that chromosome doubling in S . X townsendii, either in somatic tissue or by the fusion of unreduced gametes, produced S. anglica. This evidence can be split broadly in four categories: morphological, historical, cytological and biochemical.

B. Evidence for the Origin

1 . Morphological Evidence As noted in a previous section, when the Groves first described S. X townsendii they thought that the plant was S. alterniflora but then decided that it was S. maritima (Groves and Groves, 1879), before eventually deciding that the plant was neither of these species. The reasons for thinking initially that the plants were S . alterniflora were that the specimens were too tall to be S. maritima and that the panicles consisted of four to six spikes rather than the two usually present in S. maritima. In addition, the spikes were longer and had more spikelets than usual for S . maritima. The Groves thought, however, that these were unreliable characters for distinguishing between S . alterniflora and S . maritima. The decision to assign the forms found at Hythe to S. maritima was based on them having leaves distinctly articulated with the leaf sheaths and flag leaves that did not overtop the spikes, both characters of S . maritima. Spartina alterniflora, on the other hand, has leaves that are more continuous with the sheaths and flag leaves that always extend beyond the spikes. It can be seen from the early confusion that S. x townsendii has many characteristics in common with both S. maritima and S . alterniflora, tending towards S . alrerniflora in inflorescence structure, and towards S. maritima in foliage characters. These hybrid characteristics are borne out by the data of Marchant (1967) and Hubbard (1968). Other characteristics given as evidence of hybrid origin are the prolific rhizome production, as in S. alterniflora. and the dense tillering, as in S. maritima. The morphological evidence is, therefore, consistent with S . X townsendii being a hybrid between S. alterniflora and S . maritima. This being the case, we may ask whether the morphology of S.

A CENTURY OF EVOLUTION IN SPARTINA A N G L I C A

13

angfica is consistent with it being the product of chromosome doubling in S . x townsendii. The failure to distinguish between the two for over 50 years shows that they have many features in common. In most respects apart from height, however, S. angfica is larger than S . x townsendii (Marchant, 1967; Hubbard, 1968). Spartina anglica leaves tend to be longer and wider, the ligules are longer, and the spikes are longer as are the spikelets. This is suggestive of the so-called “gigas” morphology of increased size of plant organs resulting from increased cell size due to polyploidy. Marchant (1967) also found that the volume of pollen grains of S . angfica is twice that of S . x townsendii, a strong indication that chromosome doubling in S . X townsendii gave rise to S . angfica. Thus, the morphological data are entirely consistent with the classical theory.

2. Historical Evidence From information on the historical distribution of S . afterniflora and S. maritima it can be shown that the two species co-existed in Southampton Water immediately prior to the discovery of S . x townsendii. Spartina maritima occurs from Britain and the Low Countries in the north, through France, Spain and Portugal, and into the Mediterranean: further south the plant occurs in rather isolated populations down the west coast of Africa from Morocco to the Cape of Good Hope. Spartina maritima is usually thought of as a native of the British Isles. Throughout northern Europe, however, it shows a lack of vigour in comparison with specimens from Spain and Africa. This led Chevalier (1923) to suggest that it may only be native to Africa and was introduced into northern Europe by shipping. The first record for S . maritima in Britain is from northern Kent in 1629 (Hubbard, 1965b). Bromfield (1836) recorded it from the mouth of the River Itchen and Hythe, and Townsend (1883) described it as being “rather common” in Hampshire. By the 1900s the species had been recorded from Lincolnshire to north Kent on the east coast, and from Chichester to the Exe on the south coast (Stapf, 1908). This probably represented the greatest limit of S . maritima in Britain and by the 1930s it was becoming quite rare. Hall (1934) records that it had been “practically exterminated” in Hampshire, and in a recent survey the only location for the plant on the south coast was Hayling Island (Raybould et al., 1991b). The pattern is similar on the east coast, where since 1950 the plant appears to have been lost from Lincolnshire, Norfolk and Kent and now occurs only on high level salt marshes in Essex and southern Suffolk (Raybould et al., 1991~). There appear to be several reasons for the decline of S . maritima. On the south coast many habitats have been destroyed through land

14

A. J . GRAY E T A L .

reclamation for agriculture or industrial development; this is especially so in Southampton Water (Marchant, 1967; Raybould et a[., 1991b). On the east coast, although some populations may have been lost through land reclamation, the major factors have been erosion and successional change, the invasion of high-level S. maritima habitats by Halimione portulacoides being a particularly common cause of its decline (Raybould et al., 1991b). It has often been assumed that the spread of S. anglica has been important in the decline of S. maritima. These species rarely co-occur, however, and the spread of S. anglica and the decline of S. maritima are probably two independent indicators of ecological, sedimentary and sea level changes, rather than cause and effect. Spartina alterniflora is a native of the eastern seaboard of North America where it grows in large monospecific swards from Newfoundland in the north to the Gulf of Mexico in the south. The species is also recorded from tropical South America, although some botanists consider these records refer to the closely related species S. glabra Muhl. Sparrina alterniflora was probably introduced into Southampton Water from the USA by shipping. Bromfield (1836) suggests that the species may have been in the River Itchen since 1816. Townsend (1883) records that in 1879 it was “abundant by the Itchen from the sea upwards to beyond Southton [Southampton]”. Stapf (1908) also recorded that the plant was abundant in the Itchen and Southampton Water. The range of the species extended ultimately from Lymington in the west to the mouth of the River Meon in the east (Hall, 1934). As is the case with S. maritima, S. alterniflora has undergone a major decrease in distribution since the turn of the century. Industrial development has played a large part in this regression, although some populations were lost due to being overrun by S. anglica (Marchant, 1967). Since 1963 the species has existed at only one site, at Marchwood on the west shore of Southampton Water (Marchant and Goodman, 1969). From the records available, there appear to be two places where S. maritima and S. alterniflora occurred together, namely at the mouth of the Itchen and at Hythe. Bromfield (1836) wrote that where the Itchen entered Southampton Water “Sp. stricta grows close to patches of Sp. alrerniflora”. Townsend (1883) also recorded both species from this location. Townsend also reported S. alterniflora and S . maritimu as growing together at Hythe, and the Groves (1879) found specimens of S. X townsendii growing amongst stands of S. alterniflora. If it is assumed that S. maritima was growing at the mouth of the Itchen when S . alterniflora was first introduced, then the species could have existed together for over 50 years before the discovery of S . x rownsendii. Historical records show, therefore, that the origin of the S. anglica agg. through hybridization was feasible.

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15

3. Cytological Evidence The first work on the cytology of the S. anglica complex was by Huskins (1930a,b). Using mitotic cells from root tip meristems, he found S . maritima had 2n = 56, S. alterniflora 2n = 70 and S. anglica 2n = 126. These counts fit exactly with the theory that S . anglica is the amphidiploid derivative of the other two species. Huskins used plants of S. alterniflora from New York State rather than British material, which left him open to the criticism that he may have examined a different cytotype from the one involved in the origin of S. anglica (Marchant, 1963). Nevertheless, the results seemed so conclusive that for nearly 30 years the question of the origin of S. anglica was not re-examined. The work of Hubbard (1957) in pointing out that an F1 hybrid existed at one time probably sparked a new interest in the cytology of the S. anglica complex. Boyle and Kavanagh (1961) counted 2n = 126 for S. anglica from Ireland, although they later corrected this to 2n = 124 by the re-interpretation of multivalents (Boyle, 1973). Marchant (1963) showed that Huskins’ counts were inaccurate, and gave counts for S. maritima of 2n = 60, for S. alterniflora of 2n = 62, and for S. anglica a range from 2n = 120 to 124. He also found sterile plants at Hythe with 2n = 62 and these were designated S. x townsendii. These counts were in good, though not in perfect agreement with the classical theory. Marchant (1968) gives a detailed account of the cytology of the complex. Spartina maritima almost always forms 30 bivalents at meiosis and all mitotic cells had 2n = 60. In S . alterniflora from Britain, Marchant found some inconstancy in chromosome number in some tillers. In one tiller there was a range from 2n = 61 to 2n = 66 with a mode of 62. Meiotic counts were always n = 31 with a maximum of two multivalents per cell, suggesting a polysomic constitution of 6x + 2. Marchant (1970) also counted S. alterniflora material from North America and found 2n = 62 in all cases. This ended speculation that S. alterniflora existed as two chromosome races of 2n = 8x = 56 and 2n = lox = 70 (until the 1960s it was thought that the genus Spartina had a basic number of 7 rather than 10). In S. x townsendii Marchant (1968) found some variation in chromosome number within tillers, with a mode of 62. As with S . alterniflora, he suggested that the variation had a physiological basis as plants that had weak growth had most variation. Meiosis was shown to be complex with numerous univalents and multivalents at metaphase I (Fig. 3 ) . Trivalents were the most common multiples, although quadrivalents and other arrangements occurred, and usually less than half of the chromosomes were left unpaired. Chiasma frequencies ranged from 0.53 to 0.75

16

A. J . GRAY E T A L .

Fig. 3. Metaphase 1 at meiosis in Spartinu = 62).

X

townsendii ( 2 n = 6111

+ 1511+

per paired chromosome, and from 17.4 to 31.3 per cell, depending on univalent frequency. The highest pairing was in plants with the “dwarf brown” phenotype first described from the Dovey Estuary and the Bristol Channel by Chater and Jones (1951). Marchant found a range of chromosome numbers in S. anglica from 2n = 120 - 127, with the great majority of clones having 120, 122 or 124. All clones showed almost regular meiotic pairing (Fig. 4) with a high chiasma frequency of 0.74-0.86 per paired chromosome. Univalents occurred up to a maximum of ten per cell in aneusomatous plants due to some chromosomes having no homologue with which to pair. Quadrivalents occurred up to a maximum of five per cell, and trivalents to a lesser extent. Marchant also described possible backcrosses of S. anglica to a parental species. Two unusual clones were found in Southampton Water. One clone gave a count of 2n = ca 90 at meiosis, with a

A CENTURY OF EVOLUTION IN SPARTINA ANGLICA

Fig. 4. Diakinesis at meoisis in S. anglica (2n = 61fI

17

+ 2 , = 124).

maximum of 16 univalents and only rare multivalents. There was relatively little lagging at anaphase, and the second division was more or less regular, giving tetrads that appeared normal. Seed set in the clone was described as “abundant”. The other clone gave counts of 2n = ca 76 from mitotic cells. At meiosis up to 19 univalents were found with few multivalents. At anaphase laggards were common with as many as 11 univalents being excluded from daughter nuclei. Tetrads were irregular but seed was set occasionally. Both these clones resembled S. alterniflora in morphology (Marchant, 1967) and possibly represent a backcross of S. anglica to this parent. The genome relationships elucidated by Marchant’s work are shown in Fig. 5 , and strongly support the theory of the origin as outlined previously. Despite the very strong evidence, from several sources, some doubts still remained over the origin of S. a n g b . First, it had not proved possible to resynthesize the species from the putative parents. Marchant (1964) attempted to cross S. maritima and S. alterniflora by bagging

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A. J . G R A Y E T A L .

I I

S. martfirnu x 2n = 6x = 6 0

Salfernifforu 2n = 6x

+ 2 =62

S. x rownsendii 2n =ca 6x = 6 2

2 nd backcross 2n = ca 7 6

chromosome doubling

S onqlica 2 n = c a 12x=120,122 8 124-

agamospermy

J

polyhaploids

1st backcross 2n=cu 90

chromosome loss

\ oneuploids

Fig. 5. Genome relationships in Sparfina (based on Marchant, 1968).

flowering heads together and by placing pots in which the species were growing adjacent to each other. In neither set of experiments was seed set on either species. Marchant (1964) also used colchicine in an attempt to double the chromosome number of S. x rownsendii to produce S . anglica. Treatments included immersion of tiller and rhizome apices in colchicine solution, application of colchicine-containing agar blocks to growing points, and injection of colchicine solution inside the leaf sheath enclosing the flowering spikes. In no instance was chromosome doubling induced. Also, Raybould (1989) found that the pollen of both S. alrerniflora and S . maritirna grown in a glasshouse was completely sterile on the basis of the fluoroscein diacetate test (Heslop-Harrison et al., 1984), so that experiments to resynthesize S. x rownsendii could not be performed. A second doubt was raised by the revised chromosome counts presented by Marchant. The counts of Huskins meant the possibility that S. anglica could have arisen directly from either S. alterniflora or S. rnarifima through autopolyploidy could be discounted. The new counts made by Marchant. in which the chromosome numbers of both S. maritirna and S. alterniflora are roughly half those of S. anglica, meant that autopolyploidy must be considered.

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The available evidence was largely against an autopolyploid origin. First, the karyotype of S. anglica appeared to be a combination of the karyotypes of S. maritirna and S . alterniflora (Marchant, 1968). The chromosomes of Spartina are, however, very small and Marchant could not unequivocally confirm an allopolyploid origin on the basis of chromosome morphology. Secondly, the chiasma frequency in both putative parents is high (Marchant, 1968) and thus an autopolyploid produced from either of these species might be expected to have a high frequency of multivalents. Spartina anglica forms multivalents rarely (see Fig. 4, for example) and so the case for alloployploidy seemed strong. There are several cases documented, however, where autopolyploids form only bivalents. This phenomenon has been observed, for example, in Lotus corniculatus (Dawson, 1941), Tolmeia menziesii (Soltis and Soltis, 1988), the fern genera Asplenium and Adianturn (Lovis, 1964; Vida, 1970; Manton et al., 1970) and in several artificially produced autotetraploid crops (Timmis and Rees, 1971; Armstrong, 1971; de Wet and Harlan, 1972). Thus, autopolyploidy could not be rejected completely.

4. Biochemical Evidence The evidence that finally confirmed the allopolyploid origin came from isozyme electrophoresis. In newly formed allopolyploids the component diploid genomes are often expressed entirely (Gottlieb, 1982). Thus an allopolyploid species would have isozyme phenotypes that are the product of the addition of parental polypeptide subunits, resulting in all the parental bands being present, plus possible new hybrid bands in multimeric enzymes (Roose and Gottlieb, 1976). In newly formed autotetraploids it is assumed that no bands could be present that were not present in the parental species, and this is borne out by the evidence available (Crawford and Smith, 1984; Soltis and Soltis, 1988). It has recently been shown (Gray 1986; Raybould 1987, 1989; Gray et a f . , 1990a; Raybould et al., 1991b) that British material of S. maritima and S. alterniflora possess unique bands in several isozyme systems and that S . x townsendii contains all of these bands (see Fig. 6). In addition, S. x townsendii showed several novel hybrid bands in systems where multimeric enzymes are formed. In all systems S. anglica had phenotypes identical to S . X townsendii. This evidence completely confirmed the allopolyploid origin of S. anglica. Guenegou et al. (1988) found similar results using French material, although certain features of this work suggest that the genotypes of the parental material they examined were not those involved in the origin of S. anglica. For example, in superoxide dismutase they found that both S. alterniflora and S.

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A . J . GRAY E T A L .

Fig. 6. Isozyme phenotypes in Spartina demonstrating that S . anglica phenotypes are those expected if the species is an allopolyploid derivative of S. alterniflora and S . maritima. a (Upper): Esterase (1-4 S. maritima, 5-6 S. x townsendii, 7-8 S. anglica, 9-11 S. alterniflora, 12 S . glabra) b. (Lower): Phosphoglucose isomerase (1-3 S. alterniflora, 4-9 S . anglica, (10-13 S . maritima) (with permission from Raybould el al., 1991a).

maritima had bands that were not present in S . anglica, whereas Raybould et al. found that in British material of the parental species all of the parental bands in this system are found in the hybrids. A similar position occurred with acid phosphatases. Guenegou et al. also found esterase phenotypes with non-parental bands, which, since most esterase enzymes are monomeric (Gottlieb, 1981), further suggests that their material was not involved in the origin of S. anglica.

C . ‘Spartina

X

neyrautii”

The evidence reviewed in the preceeding sections confirms beyond reasonable doubt the nature of the origin of S . anglica. There is, however, further evidence from a plant known as S. x neyrautii. In south west France there was also an accidental introduction of S. alterniflora. It was discovered by Loiseleur in the Ardour Estuary near Bayonne in 1803. By the early 1900s it had spread along the coast for about 25 miles from Capbreton to the Bidassoa Estuary (Stapf, 1908). S.

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maritima is native to this area (Mobberley, 1956). In 1892, Neyraut collected an unusual form of Spartina from Hendaye (Mobberley, 1956) and the status of this plant and similar types found in the area caused much confusion. When Neyraut discovered the plant he though it to be identical to S. x townsendii (Chevalier, 1923). In 1894 Foucaud described it as a hybrid between S. maritima and S. afterniflora (Stapf, 1908), but named it S. x neyrautii as he believed it had certain differences from S. x townsendii. The hybrid nature of S. x neyrautii was confirmed by Marchant (1977), who found that it had the same chromosome number and meiotic behaviour as S. X townsendii and also very similar morphology. Raybould et af. (1991a) confirmed that the plant is a S. maritima x S. afterniflora hybrid using isozyme phenotypes, and showed that S. X neyrautii must be considered as a synonym of S. X townsendii. The fact that S. X townsendii has been discovered in the only two places where the distributions of S. afterniflora and S. maritima are known to have overlapped is a further confirmation of the nature of its origin.

D. Problems Over the Origin Two questions remain to be answered over the origins of the S. angfica agg. First, did S. angfica sensu strict0 arise more than once, and second, do the two independent origins of S. X townsendii represent reciprocal hybridizations, as suggested by Arber (1934). The first question has been addressed by Raybould et al., (1991a). They found that S. angfica has extremely low levels of genetic variation at isozyme loci (see below), suggesting a single origin. The parental species and S. x townsendii, however, also show little variation and so a multiple origin from uniform parents cannot be discounted. Further research using more sensitive methods of detecting genetic variation (such as “genetic fingerprinting”) may resolve this problem. There is no available information on the second question. A possible solution, however, may be a study of the inheritance of chloroplast or mitochondria1 markers, which have been used recently to demonstrate reciprocal origins for Tragopogon species in the USA (Soltis and Soltis, 1989).

V. VARIATION IN SPARTINA ANGLICA

A. Sources of Genetic Variation in Spartina anglica The available evidence on the levels of genetic variation in British populations of the parental species of S. angfica is limited to an isozyme

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survey by Raybould et al. (1991b). This found that the single extant population of S. alterniflora was monomorphic, and, although there was some evidence of variation in S. maritima, British material of this species also appeared to be largely monomorphic with isozyme variation limited to two comparatively rare forms of the enzymes shikimic acid dehydrogenase and glutamate oxalacetate transaminase. As indicated earlier, this lack of significant levels of isozyme polymorphism in the parental species means that it is not possible to distinguish between single or multiple occurrences of either the original hybridization event which produced S. X townsendii or the subsequent chromosome doubling which produced S. anglica (as has been done with other allopolyploid species such as Senecio cambrensis (Abbott and Ashton, 1989), Tragopogon (Roose and Gottleib, 1976) and Pfagiomnium medium (Wyatt et al., 1988)). However, the lack of enzyme polymorphism in the parental species does at least suggest that even if multiple events have occurred, they are likely to have resulted in genetically identical forms of S. anglica. Since the original formation of the species, the generation of variation in S. anglica may have occurred in a number of ways apart from conventional mutation. One possible source of variation is through backcrossing and subsequent introgression with one of the parental species. Some evidence of the type of chromosomal polymorphism that might be associated with such events exists (see Section IV above). Two further potential sources of variation are chromosomal in origin. In an allopolyploid species such as S. anglica, in which chromosome doubling has restored fertility, pairing at metaphase I of meiosis often only occurs between homologous chromosomes, i.e. equivalent chromosomes from the same parental genome. This phenomenon, known as preferential pairing, has been extensively studied in polyploids such as wheat (e.g. Riley and Chapman, 1958). Strict preferential pairing maintains the variation between the parental genomes in the form of fixed heterozygosity (e.g. Aung and Evans, 1987). In contrast, if homeologous pairing. i .e. pairing between the equivalent chromosome from different parental genomes, occurs, recombination between the parental genomes is possible. Homeologous pairing- even at relatively low frequenciescould lead to the gradual release of genetic variation from recombine? tion between the parental genomes. A second chromosomal source of variation is through aneuploidy (i.e. the gain or loss of chromosomes) as the result of either unbalanced segregation at meiosis or somatic events. In general, plant species, especially high polyploids such as S. anglica, are relatively tolerant of the resulting unbalanced chromosome dosage that results from aneuploidy, at least through the vegetative stages of their life cycle. However,

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23

the resulting complications at meiosis normally lead to at least some level of reduced sexual fertility. This is not a problem in species which, like S. anglica, propagate extensively by asexual means (Gibbs et al., 1978).

B. Evidence for Genetic Variation in Spartina anglica 1. Morphological Variation Although many authors have reported field observations of morphological variation in S . anglica, it has only recently been systematically studied. Some evidence of phenotypic variation comes from piants which appear “distinct” under field conditions, for example the “brown dwarf” type identified by Chater and Jones (1951). Marchant (1968) has shown that many dwarf forms have a chromosome number of 2n = 62, suggesting that they have a polyhaploid origin. Some other “distinct” types have been shown to revert to typical S. anglica morphology when brought into artificial cultivation (Raybould, 1989). Spartina anglica has also been observed to display zonal variation in morphology across some salt marshes. For example, Marks and Mullins (1984) and Marks and Truscott (1985) have reported the zonation of S. anglica on the salt marshes of the southern shore of the Ribble estuary, in north west England. They recognized four zones on the basis of shoot density and vegetative vigour. Zonation of S. anglica has also been extensively studied on the salt marshes of the north shore of the Dee estuary in Cheshire. Taylor and Burrows (1968) identified differences in time of flowering, seed production, rates of tillering and number of overwintering shoots across this marsh. If these zonal differences are genetic in origin and the result of adaptive differentiation in response to different environmentally induced selection pressures across the marsh, extremely rapid microevolution must have occurred since the original colonization of the Cheshire marshes on the Dee in 1945. However, genetic variation cannot be inferred from such observed zonal differentiation alone since any genetic effects will be confounded with age effects and a range of environmental effects. Material from the Dee estuary salt marsh has been subjected to careful study in a series of common garden experiments. Plants derived from single tillers, from distinct zones showed significant differences even after 2 years of cultivation under common environmental conditions (Hill, 1990). In addition, plants transplanted into the zones from which they originated showed significantly greater rates of survival than plants transplanted between zones. This appears to indicate that there may be a genetic component to the observed zonal differentiation across the Dee estuary salt marsh. However, the unambiguous resolution of

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such morphological variation into its genetic and non-genetic components in such a phenotypically plastic species as S. anglica poses a number of major experimental problems. Since individual clones may be long-lived, there is considerable potential for carry-over effects into clonal trials under common garden conditions. Carry-over effects may arise in several ways. For example, differences in the nutritional status of the clones can produce effects, as in the differences in salt tolerance among clonal lines of Puccinellia maritima which were found to be caused by variation in the salt concentration of the soils from which the clones were sampled rather than by differences in genotype (Gray and Scott, 1977). Carry-over effects may also occur in perennials from differences in clone age, older ramets being less able to change their phenotypes in response to environmental change (Breese et al., 1965). Different levels of viral infection, to which long-lived clonal plants are especially prone, may also lead to non-genetic differences in gross morphology, yield and competitive ability (Silander, 1985)-although its effects on growth are unknown, a leaf mottling virus, spartina mottle virus, was identified by Jones (1980) and isolated from plants in several populations. Unfortunately, the differences in morphological variation in S. anglica populations, both zonally in the Dee and Ribble estuaries, and within and between complementary zones on different salt marshes (Thompson, 1990), are largely of the type which could arise from differences in clonal age or viral infection levels. Although there is no direct evidence of such effects, plants from the lower zones, which are usually younger, are generally more vigorous and have greater phenotypic plasticity (Thompson, 1989, 1990; Thompson et al., 1990, 1991 a,b,c. Thompson et al. (1990, 1991a,b,c), on the basis of further common garden, reciprocal transplant and genotype/environment experiments, suggest that most of the variation is due to age-related decline in vigour, and thus to somatic rather than genetic differences between populations. The only way to resolve this variation into its genetic and environmental components would be to carry out an appropriately designed trial of seed progenies (using the type of experimental designs summarized in Lawrence (1984)). This has not proved possible in S. anglica because of the unpredictable pattern of seed set in recent seasons, the problems of controlled germination of seed, and the difficulties of growing such a large, vigorous, rhizomatous salt marsh perennial grass in a fully randomized field trial. Ideally, seed progenies produced under glasshouse conditions should be used, to avoid environmental differences during seed maturation, which can cause carry-over effects (e.g. Funk et al., 1962; Nelson et al., 1970). Even strong short-term environmental stimuli at the seedling stage can influence the phenotype of that plant’s seed progeny (Hague and Jones, 1987).

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25

2. Biochemical Variation Although high levels of seed protein variation and extensive isozyme polymorphism are features of many higher plant populations (e.g. Gepts, 1990; Hamrick and Godt, 1990), an extensive study of protein variation in S. anglica conducted by Raybould et al. (1990b) found no evidence of seed protein variation, and found only a single GOT isozyme variant in a survey of 12 different enzyme systems. In contrast, the polyploid grass P. rnaritirna was found to be highly polymorphic for every isozyme system studied in the population coexisting with monomorphic S. anglica on the Dee estuary salt marshes (Raybould et al., 1991a) (Fig.7), and in several other populations (Gray et al., 1979). It

Fig.7. Comparison of the variation in esterase isozymes in clones of (a) (Upper): S. angfica and (b) (Lower): P. rnaritirna collected along the same 300 m transect on the Dee estuary (with permission from Raybould et af., 1991a).

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appears, therefore, that at the protein level S. anglica has extremely low levels of genetic variation. The single G O T variant identified by Raybould et al. (1991a) was characterized by the loss of several G O T bands which correlated with a small reduction in chromosome number, suggesting an aneuploid origin. The variant appeared to be largely confined to Poole Harbour and showed considerable variation in frequency between creeks. Its sporadic distribution elsewhere may reflect either a series of sampling events, as Poole material has been distributed throughout Britain, or alternatively a series of independent aneuploid origins. Of course, a lack of electrophoretically detectable variation does not conclusively signify an absence of genetic variation. Isozyme markers may show lower levels of variation than quantitative traits, as in Hordeum mitrinum (Giles, 1984) and Xanthium strumarium (Moran and Marshall, 1978; Moran et al., 1981). In general, however, seed proteins tend to be more variable than isozymes, probably because of fewer functional constraints on seed protein structures (Gillespie and Blagrove, 1975; Righetti et al., 1977). For example, Doll and Brown (1979) estimated that hordeins were 10-30 times as variable as isozymes in barley. The completely invariate protein profiles of S. anglica seeds are therefore strong indicators of genetic uniformity in the species-particularly as we assume the seeds to be produced sexually, whereas common garden trials use vegetative tillers which could be entirely clonal. The general conclusion that can be drawn from the existing information is that, apart from the sporadic occurrence of chromosome variation resulting from aneuploidy, polyhaploidy and back-crossing (see above), there is little unambiguous evidence of genetic variation in S. anglica. This is what one might expect from both its origin and the pattern of its subsequent spread, since British material of both parental species shows little evidence of variation; the original hybridization event and the subsequent chromosome doubling appear to have occurred at very low frequency, possibly only once, and the process of small-scale clonal sampling and subsequent multiplication which has been used to plant out new sites would have dramatically reduced any variation which might have been present. Finally, conventional mutation is unlikely to have led to the generation of significant levels of genetic variation over such a short evolutionary timescale.

3. Genetic Variation in Other Clonal Salt Marsh Grasses The lack of genetic variation in S. anglica which is likely to be a major factor in the future evolutionary success of the species, is in contrast to at least some other clonal salt marsh grasses. Genetic differentiation, some of adaptive significance, has been demonstrated in P. maritima

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using clonal and seed progeny trials and isozyme studies (Gray, 1985a, Gray er al., 1979, Gray and Scott, 1980), as has genetic variation in salt tolerance in salt marsh Festuca rcibra using clonal and seed progeny material (Rhebergen and Nelisson, 1985). Perhaps more significantly, Silander (1984, 1985; Silander and Antonovics, 1979) provided evidence for adaptive genetic variation in North Carolina populations of the related Spartina patens, which, although based on trials of clonal material, was supported by isozyme variation. Indeed, clonal plants generally are found to possess genetic variation, most having multiclonal populations with most clones restricted to one or a few populations and rarely being widespread (Ellstrand and Roose, 1987).

VI. THE ECOLOGY OF SPARTINA ANGLICA

A. Rates and Pattern of Spread and Dispersal As detailed above, S. anglica was successfully introduced to coastal and estuarine mudflats throughout the world, especially during the 1920s and 1930s. Natural dispersal of seed, usually in entire spikelets, over long distances may have occurred by tidal currents, shipping, or on the feet of wading birds (as suggested by Hardaker (1942) as the means by which it reached an inland site on the Droitwich Canal (Goodman et al., 1969)). The rate of spread at individual sites has varied greatly but in many areas has involved the rapid spread from transplants or other propagating units to form a continuous sward. This process has acquired a terminology (Hubbard, 1965a), young plants derived from seedlings or plant fragments expanding into more or less circular “tussocks” which fuse to form “clumps”, often of irregular outline, with clumps finally coalescing to form a “sward”. Caldwell (1957) noted a pattern of concentric rings of different shoot density distribution in expanding tussocks (which she termed “auxoclones”), growing without competition on open mudflats. The rings are produced annually by the peripheral growth of rhizomes but are not detectable in all expanding tussocks. The formation of swards is not always a rapid or continuous process. Many of the early transplants failed or appeared to be held at the tussock phase. The known reasons for failure include inappropriate site conditions (too unstable, too sandy) and frost, particularly at the northern edge of the species’ range; the death of transplants by frost has been recorded from Scotland (J. Bryce, unpublished), the north Netherlands (Kamps, 1962), north Germany (Konig, 1948) and China (Chung, 1990). The very slow spread in some areas may have been due to the introduction of only S. x townsendii plants, which are both less vigorous

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vegetatively and unable to spread by seedlings. This is believed to account for the slow expansion of populations in several Australian and New Zealand marshes (Ranwell, 1967). Even where the fertile amphidiploid has been introduced and has become successfully established, the rate of spread may be remarkably discontinuous. Particularly characteristic of western and northern sites in Britain is a pattern where initial colonization, whether slow or rapid, is followed by a long period of around 20, 30 or even 40 years during which there is little or no expansion in area of the tussocks or swards. This is then followed by a sudden burst of population growth. Gray et al., (1990a) have documented such a pattern in the Conwy estuary in North Wales, where the rapid colonization of a large mudflat in the upper estuary occurred in the early 1970s, adjacent to a sward which had not grown much in area for almost 20 years. Deadman (1984) records a similar rapid expansion between the mid-1960s (the Hubbard and Stebbings (1967) estimate) and 1982 in several other Welsh estuaries to which the plant had been introduced in the 1930s. In Morecambe Bay there has been rapid colonization in recent years of parts of the River Kent estuary, where only isolated clumps were present up to 1982 (R. Scott, pers. comm.; Whiteside, 1984). Further north, in the Cromarty Firth (57.5 O N ) , the populations at Dingwall Bay, planted in 1932, had not increased much in size by 1955 but were expanding by around 2 m per year (with a maximum of 8 m in 1976) during the 1970s (Smith, 1982). Similarly, at nearby Udale Bay, populations established in 1948 expanded in area from 1358m’ in 1970 to 4228m’ a decade later (Smith, 1982). The causes of this pattern are unknown, but evidence from several sites indicates that sudden spread is marked by a successful phase of seedling establishment, lasting only 1 or 2 years, followed by expansion of the more-or-less evenly aged tussocks to eventually form a sward. These events require the coincidence of high seed production and suitable conditions both for seedling establishment and subsequent tussock expansion. It is not clear which of these conditions is most frequently absent in sites where S. anglica is failing to spread. Although the species is noted for the unpredictable production, viability and germination of its seeds (Hubbard, 1970; and below), it is equally clear that in some areas changes in the sedimentary regime have preceded the expansion of the population. For example, the tidal flats recently colonized in Morecambe Bay became notably muddier, due to a change in position of the River Kent low-water channel, before they were invaded. In some areas, colonization of the tidal flats below an existing sward occurs by the successful establishment of plant fragments. This may

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follow uprooting by storms, as at Bridgwater Bay (Ranwell, 1964a), or by grazing animals, as at Udale Bay (Smith, 1982). Both plant fragments and seedlings may occur as propagating units in the same area (Chater and Jones, 1957; Taylor and Burrows, 1968). Under suitable conditions, seedlings may occur at densities up to 13000 m-’ on bare mud and up to 9750m-’ in the sward, where most die, the sward being maintained by rhizome formation and tillering (Goodman, 1960; Goodman et al., 1969). The factors affecting seedling establishment on bare mud have been studied experimentally by Groenendijk (1986), who demonstrated a critical interaction between mudflat elevation, and hence hydrodynamic stability, and seed burial depth in determining seedling emergence and growth. Seeds buried between 1 and 3 c m had the greatest chance of establishing as seedlings, those nearer the surface being lost through desiccation or sediment movement, and deeper buried seeds deteriorating rapidly in viability. The lower elevational limits of S. anglica in Groenendijk’s study area (the Oosterschelde in the south west Netherlands) were controlled mainly by wave action uprooting seedlings, supporting Morley’s (1973) suggestion that this is a major limiting factor. Seed production in S. anglica is extremely variable, both temporally and spatially. Goodman er al., (1969) recorded 92% seed-set at Lymington in 1954, followed by 18% the next year. Taylor and Burrows (1968) found both seasonal (15%, 38% and 64% in successive years) and tussock to tussock variation in the Dee estuary. Detailed studies of the Ribble estuary marshes have revealed zonal variation in filled spikelet production, Marks and Truscott (1985) recording in 1978 2.3%, 77.1%, 16.9% and 3.7% in successive zones from the pioneer to the upper limit of S. anglica. However, less than 5% of the filled spikelets produced seed which germinated. Mullins and Marks (1987) recorded values ranging from 3.1 to 6.8%, from 6.7 to 25.8%, from 2.1 to 4.5% and from 4.0 to 5.7% over three years in the comparable zones of a nearby marsh. Although filled spikelet production in Poole Harbour populations averaged 27% in 1984, no seed germinated, and viable seed production has been exceptionally low in every year since 1983 (Gray el a f . , 1990b). It is not surprising that seed does not set in most years since seed production may involve the breakdown of a self-incompatibility system. Raybould (1989) used the petri-dish method of Lundqvist (1961) to examine cross-pollination of S. anglica from geographically widely separated populations. In all cases, growth of the pollen was arrested rapidly and the pollen tubes had the typical curled appearance found in a “weak” self-incompatibility reaction (sensu Shivanna et al. , 1982). Grasses have a two-locus self-incompatibility system, and in polyploid grasses only one allele needs to be matched at each locus in the pollen

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and stigma for a self-incompatibility reaction to result (Fearon el ul., 1984a,b,c). As S. anglica is a duodecaploid with a very narrow genetic base, it is extremely unlikely, that there are any cross-compatible genotypes within the species. When seed is produced, it is only in inflorescences which appear relatively early in the year. Inflorescences emerging after August on the Ribble marshes all failed to produce mature filled seed, irrespective of their position on the marsh (Mullins and Marks, 1987; Marks and Mullins, 1984). The delayed development of inflorescences in the higher parts of the marsh was thus a major factor in the much lower proportion of seed produced in this zone. The conditions under which the self-incompatibility system may break down, sporadically producing large numbers of viable seed, are unknown. The occasional production of seed in glasshouse-cultivated plants suggests that higher than average temperatures and humidity may be a factor. High temperatures are known to lead to a breakdown in self-incompatibility in other species, for example Trifolium repens (Chen and Gibson, 1973) and Lilium longiflorum (Ascher and Peloquin, 1966). Additionally, Stapf (1913) noted that high seed set occurred following a hot, dry summer in 1911. In summary, the natural spread of S. anglica populations characteristically comprises two phases: the initial invasion and establishment of seedlings on open mudflats. and the subsequent radial expansion of tussocks to form clumps and ultimately a sward. Seedling establishment is a rare event in most populations and often follows many years of purely clonal expansion. This may be when mudflat conditions reach a threshhold of sediment type or stability or following a year in which, unusually, there is high seed production (or when both circumstances coincide). In the 1970s and 1980s there appears to have been a sudden expansion of several populations in estuaries in the west and north of Britain to which the plant may have been introduced in the 1930s and 1940s. Of the factors related to tidal submergence which may limit the seaward spread of populations, the tidal uprooting of seedlings is known to be particularly effective.

B. Growth and Production The annual cycle of growth in S. anglica comprises the production in leaf axils in November of overwintering buds which then grow fast and flower in the long days of summer and early autumn, the flowering culms dying as a new generation of buds is formed the following November (Goodman et al., 1969). Rhizomes develop during the winter, in response to short days (Goodman 1960). Flowering extends from July to November and even, in mild winters, into the following year.

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Spartina anglica swards occur across a wide range of sediments ranging from clays, fine silts and organic muds to more sandy substrata and even shingle where it is regularly inundated by the tide. They also experience a wide range of sediment accretion rates, generally varying from 3 mm year-' to as much as 8-10 cm year-' (Bird and Ranwell, 1964; Ranwell, 1964a; Lee and Partridge, 1983), but extremely high annual rates recorded for short periods in actively growing swards include 17 cm at Bridgwater Bay, England (Ranwell, 1964a), 20.cm l,at Sloedam, the Netherlands (Verhoeven, 1951), and 24 cm and 26 cm at Qidong and Rudong, respectively, in China (Chung, 1990). This range of conditions is reflected in the variation in productivity. Estimates of above-ground standing crop from British marshes include an average of around 825 gm-2 (dry weight) on high marshes and a range from 418 gm-2 to 1232 gm-' at Bridgwater Bay (Ranwell, 1961, 1964b), an average of 1145 gm-2 (Dunn et a f . , 198l), from the Ribble estuary marshes, and a range of values from 400-500 gm-' for Seafield Bay in Essex (Long et al., 1990). Estimates of above-ground net primary production range from 475-700 gm-' year-' at Seafield Bay to 1600-1850 gm-2 year-' in the Ribble (Scott et al., 1990; Scott in Marks and Truscott, 1985). Groenendijk (1984) reported an annual aboveground net primary production of 1162-1649gm-2 in a Dutch salt marsh. All these figures are considerably lower than those reported for S. alterniflora marshes in North America (e.g. Turner, 1976) where maximum values of 4 800 g me2 of annual above-ground net productivity may be reached in tall creek-edge stands (Gallagher et al., 1980). In one of very few studies of below-ground biomass in S. anglica, Dunn (1981) and Long et al., (1990) have shown that, although there is much year-to-year variation, on average 75% of the biomass was below ground, with rhizomes accounting for around half the below-ground total, and that peak below-ground biomass was reached in the early winter. They also estimated that gross primary production, which takes account of the turnover of shoots, roots and rhizomes between samples, in the Seafield Bay marsh was around 4500gm-2year-', some three times the estimate of net production. Perhaps the most interesting aspect of the growth of S. anglica is the late annual development and temperature dependence which stem essentially from its utilization of the C4 photosynthetic pathway. The species is one of a small number of C4 species in the British flora (eight known, of which at least three are introductions-Long (1983)) and is exceptional among them in being a dominant component of large areas of vegetation (Long et al., 1975). The C4 species, in which the first product of photosynthetic C 0 2 fixation is oxaloacetate, in contrast to phosphoglycerate in C3 species, mostly occur in tropical and subtropical regions and are rare in cool, temperate

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climates. The "Kranz" leaf anatomy (numerous chloroplasts in the bundle sheath cells) and low C 0 2 compensation point (the ambient concentration of C 0 2 at which the net flux of C 0 2 at the leaf surface in the light is zero) typically found in C4 species was reported for S . anglica in 1975 (seemingly independently by Long er al. (1975) and Mallott et al. (1975)). Further work by S. P. Long and his colleagues has established that S. anglica shows at least partial adaptation to cooler climates, being able to tolerate, and maintain photosynthesis at, lower air temperatures than nearly all other C4 species (Dunn et al., 1981; Long and Woolhouse, 1978; Long and Incoll; 1979, Long, 1983; Thomas and Long 1978). Although at optimum temperatures C4 species exhibit greater nitrogen and water-use efficiency than comparable C3 species, a factor which may be important in the high tolerance of salinity in S. anglica and its relatives, their photosynthetic rates are generally inferior at 1 6 ° C and at 7-9°C photosynthetic C 0 2 assimilation ceases. In S. anglica, however, photosynthetic rates of individual leaves equal those of the C3 species P. maritima at 5 "C and 10 "C and exceed them above 1O"C, reaching to 50% higher at 25 "C. (Similar results have been obtained from one of the other temperate perennial C4 species, Cyperus longus (Jones et al., 1981.)) Despite this partial adaptation, demonstrated in the laboratory in plants grown at temperatures above 12 "C, detailed field studies in south east England have shown that, whilst some leaf growth may occur throughout the year, significant canopy development does not begin until the mean air temperature exceeds 9 "C (Dunn, 1981; Dunn et al., 1981; Long, 1983). This is revealed in the contrast in growth pattern with P. maririma shown in Fig. 8. P. maritima shows an increase in shoot weight in March, when air temperatures rise above 5 "C, growth peaking in June and July, whereas S. anglica does not show an increase until June, when temperatures exceed 9 "C, and peaks in October. Despite this, both species have a similar net primary production, that in S . anglica being produced over a shorter and later period (Hussey and Long, 1982). Indeed, Long (1983) concludes that adaptation of S. anglica, and other C4 perennials, to cool climates has been through their ability to grow and reproduce during the relatively short part of the year when average temperatures rise above 9-10"C, as well as the partial adaptation of C02-assimilatory capacity to low temperatures. This feature of the growth of S. anglica may be reflected in the zonal pattern of seed production in more northerly populations, described above, where the later flowering of the high marsh plants was attributed by Mullins and Marks (1987) to the delay of tiller development by depressed spring and early summer soil temperatures in this zone (Marks and Mullins, 1990). Detailed studies of the tiller dynamics of the

33

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I

I

M

1

A

I

M

1

J

1

J

1977

1

A

1

S

1

O

1

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1

O

I

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I

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M

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Fig. 8. The dry weights of living shoot material in monotypic stands of S. anglica and P. mnritima in northeast Essex; also shown is mean air temperature. Modified from Long (1983).

Dee estuary population lend support to this idea. Taylor and Burrows (1968) report earlier and more vigorous shoot production in lower marsh populations, as do Hill (1990), Thompson (1989), and Thompson et al. (1991a,b), who show that shoot production also continues later in the pioneer zone and that there are clear zonal differences in tiller demography. The more vigorous growth in the pioneer zone is marked by higher tiller production and the greater proportion of tillers which flower within a year. Thompson el al. (1991a) also reports markedly later flowering in a population of S. anglica in the Cromarty Firth, a similar trend of later flowering in northern populations having been

34

A . J . G R A Y ET AL.

observed for S. alterniflora on the east coast of the USA (Somers and Grant, 1981). The contrast in seasonal growth and production between S. anglica and P . rnaririrna consequent upon their utilization of different photosynthetic pathways suggests that the outcome of their competitive interaction will be strongly influenced by climate, and particularly spring and early summer temperatures and the length of the growing season. Although confounded by variation in sediment type and grazing, there is evidence that the ability of P . rnaririrria to replace S. anglica on the northern Dutch marshes of the Waddensea area is related to its relatively early seasonal development, and thus shading of young S. anglica shoots (Scholten and Rozema, 1990). In the south Netherlands Delta area, dense monospecific swards of S. anglica are currently persisting or, as in southern Britain, are only slowly being invaded by P. maririrna and other species. The rate at which S. anglica is being replaced by P . niaritirna and other species, notably Aster tripoliurn and Suaeda rnaritima, in the north Ribble marshes (Hill, 1987), and the decreasing upper limit to S. ariglica swards with increasing latitude (described below), are also indicators of climatic effects on the species' growth and competitive ability. Recently, Long (1990; Long er al., 1990) has examined the intriguing consequences of projected climatic changes, due to the accumulation of so-called greenhouse gases, on the primary productivity of S. anglica and P. rnaritirna using a simple analytical model. The model, which predicts primary production from light interception and conversion efficiencies, suggests that production in both species will increase, but for different reasons. The C 3 species may be expected to increase primary production principally by increased conversion efficiency in a high C 0 2 environment through reduced photorespiratory losses. However, of greatest significance in the C4 species is the effect of elevated temperatures which would enable it to increase early spring growth (reaching the point where leaf area index is sufficient to intercept 30% of the incoming radiation 50 days earlier with a 3 "C temperature increase), and thus the size and photosynthetic capacity of the canopy. This gives a predicted increase in annual production from 1.3 kgm-2 in 1978 to 2.1 kgm-' in 2050 (Long, 1990). Although produced relatively late and over a short period, the high and uniformly dense biomass presented by S. anglica swards has attracted various forms of exploitation. These include cropping for silage (Hubbard and Ranwell, 1966), paper-making (Chung, 1990), and, more recently, as a potential biofuel crop (Scott et al., 1990). Although Bryce (1941) reported no diminution in yield during 3 successive years' cropping of an English south coast marsh. Scott er al. (1990), again

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35

working on the Ribble marshes, demonstrated a decline in yield from 16 dry tonnes ha-’ year-’ to 8 tonnes ha-’ year-’ after 3 years in autumnand winter-harvested plots. Harvesting plots early in the season resulted in their colonization by P. maririrvia, whilst cuts in January maintained S. anglica cover (by increased stem density). The production of a P. maritima-dominated salt marsh by repeated summer and early autumn croppings on the Ribble parallels the effects of grazing, there and on other marshes. Puccinellia maritima replaces S . anglica at a greater rate in marshes grazed by sheep, the rate depending on local site conditions (Ranwell, 1961, 1964b, 1967).

C. The Niche of Spartina anglica The establishment of S . anglica swards on mudflats throughout northern Europe has led to the view that the species is phytosociologically in some way out of equilibrium with or is a “disharmonic element” (Beeftink, 1975, p. 431) in its environment, both by blocking the “natural” succession to other communities and exhibiting extensive die-back in the longer established sites (Beeftink, 1975, 1977). As indicated above, however, examples of succession from S . anglica swards are fairly widespread and are increasingly being reported from studies of permanent plots (Hill, 1987). The replacement of S . anglica by P. maririma was reported from the French coast as early as 1926 (Oliver, 1926). Ranwell’s (1964b) studies at Bridgwater Bay charted the invasion of swards by (usually taller) species such as Scirpus maritimus, Phragmites australis and Elymus pycnanthus (= Agropyron pungens). Under brackish conditions, S. anglica may be succeeded by Juncus maritimus and F. rubra (Packham and Liddle, 1970), and even in the south coast sites where it is dying back in the lower zones, the upper limit of the species may be invaded by P. maritima and Halimione portulacoides (Gray and Pearson, 1984). Whether it is in the active sward-building phase or is dying back, S. anglica is now found throughout most of its range as a belt of vegetation immediately seaward of the other salt marsh communities. Although individual clumps may be found at higher elevations, particularly in low-lying areas such as creek edges or pans, and may even occur around pools within reclaimed land, the upper limit of the sward is commonly marked by a gradual but often clear transition to other perennial vegetation (Fig. 9). The boundary between S. anglica swards and the landward vegetation tends to be more obvious in areas of high tidal range and on more steeply sloping marshes (see, e.g., Plate 1 in Armstrong et al., 1985). The upper limit of S . anglica is likely to be determined by its competitive interaction with the species to landward

36

A. J . GRAY E T A L .

Fig. 9. Diagrammatic representation of the zone typically occupied by S. anglica in British salt marshes.

(listed above) under conditions which increasingly favour the competitor over S. anglica as mudflat levels rise due to sediment accretion. The factors related, to increasing tidal submergence which appear to determine the lower limit of individual salt marsh species will affect both the outcome of competition between S. anglica and the species above it, and also the lower limit of S . anglica itself (Gray, 1985a). In some areas, as demonstrated in a classic study of Salicornia species by Wiehe (1935), and by Groenendijk (1986) for S. anglica, the physical effects of tidal submergence are important in limiting the plant’s seaward spread. The effects of frequency and periodicity of tidal submergence are also likely to be important. For example, Armstrong et al. (1985) showed that, in the S. anglica zone of an English east coast marsh, reducing conditions persisted throughout much of the soil profile and were indifferent to periods of exposure, there being phases of oxidation only near the surface and at neap tide periods. By contrast in the zones above S. anglica soil redox potentials were either lowered monthly during the high spring tides or there were longer periods of oxidation. Such tide-related limiting factors were presumably important in restricting the seaward advance of pioneer species such as P. maritima, Aster tripolium and Salicornia species before the advent of S. anglica. The ability of S. anglica to colonize the zone below these species (and in

A CENTURY OF EVOLUTION IN SPARTINA ANGLICA

37

some cases to invade upwards into the Puccinellia zone) stems from the suite of evolved physiological and physical attributes of the genus conferring salinity- and flooding-tolerance. These include not only the utilization of the C4 photosynthetic pathway but also the ability to oxidize phytotoxins such as Fe(I1) compounds and sulphides (Carlson and Forrest, 1982; van Diggelen, et al., 1986) and the vigour and tolerance of sediment accretion, greater in the allopolyploid than its native progenitor. Together these attributes have enabled S. angficu to invade a zone which was formerly unoccupied, at least by perennial vegetation, and could be regarded as a vacant niche (i.e. unexploited space and resources, usually only recognized when it has been occupied-Gray, 1986). A recent attempt to quantify the niche of S. anglica in terms of physical and tide-related variables is described by Gray et al. (unpublished). Their data set comprised measurements of the upper and lower limit of S. anglica along 107 levelled line transects across salt marshes in 19 estuaries in south and west Britain from Poole Harbour to Morecambe Bay. A close relationship was observed between both the upper and lower limits and the tidal range of the estuary (Fig. 10). Indeed, simple linear regressions of distribution limits against spring tidal range indicated that 88% and 86% of the variation in upper and lower limits, respectively, could be accounted for by variation in tidal range alone. Other variables which significantly improved the ability of multiple regression equations to account for the variation in distribution limits were “fetch in the direction of the transect” (a measure of exposure), “estuary area”, “estuarineness” (the position of the transect along the gradient from estuary mouth to upstream tidal limit), and, in the case of upper limits, “latitude”. The lower limit (LL) of S. anglica (in metres OD Newlyn) was best described by the equation

LL = -0.805 + 0.366 (SR) (0.102) (0.019) ( R 2 = 93.7, S = 0.35)”

+

0.053 (F) + 0.135 (log,A) (0.016) (0.025)

where SR = spring tidal range (m), F = fetch in the direction of the transect (km) and A = estuary area (km’). Spartinu therefore extended further down the shore than would be predicted from tidal range effects alone on those transects with a shorter seaward fetch and in smaller estuaries. * In the equations above, the standard errors of the regression coefficients are given in brackets below each coefficient -the significance of the particular variable can be tested using Student’s I computed as I = b / S E ( b ) which is distributed as I with N - I - p degrees of freedom, where N = the number of transects and p = the number of regressor variables in the equation. S = Residual standard deviation in metres.

38

A . J . G R A Y ET A L .

The upper limit (UL) of Sparrinn was described by:

U L = 4.74 + 0.483 ( S R ) + 0.068 ( F ) - 0.099 ( L ) (2.29) (0.028) (0.020) (0.045) (R' = 90.2, S = 0.50) where SR =range and F = fetch, as in the equation above, and L = latitude (in degrees N expressed as a decimal). The upper limit was similarly affected by fetch but also varied significantly with latitude: the further north the marsh, the lower down the shore relatively speaking was the upper limit of the Spartina sward. Both equations accounted for more than 90% of the variation in the upper or lower limits of the species, a remarkably high proportion for a biological model. The lower limit could also be predicted in relation to its deviation from the Mean High Water Neap tides (MHWN)-tidal range again accounting for most of the variation. In general, S. anglica was able to extend below MHWN in estuaries with a spring tidal range of less than 7 m (Fig. 10). The reasons advanced for the precision with which the vertical limits of S. anglica could be predicted were threefold (Gray et al., unpublished). First, tidal range provides a good general estimator of the complex of interacting factors which are likely to be determining the limits of the species distribution, encompassing variation in turbidity and in mechanical factors such as water depth and current speed. Second, such physical factors clearly control S. anglica limits, particularly the lower limits, to a degree unusual in natural vegetation. In the absence of biological competitors, the species has extended seawards to the point where tide-related variables limit its further spread (Hubbard, 1969; Morley, 1973; Groenendijk, 1986). Significantly, the upper limit, where S. anglica may interact with a number of different species, is likely to be fixed partly by competition, and thus by biological as well as physical factors. These factors are less easy to quantify. Interestingly, the downshore movement of the upper limit which occurs with increasing latitude may reflect overall changes in competitive ability of S. anglica under lower temperatures and a shorter growing season. Third, the relatively recent evolution of the species and its apparent genetic uniformity (Raybould et al., 1990b, and above) have precluded extensive population differentiation of the type which enables species such as P . rnaritima and Aster tripolium to increase their amplitude in salt marsh habitats (Gray et al., 1979; Gray and Scott, 1980; Gray, 1985a, 1987). It is probable that the realised niche (sensu Hutchinson (1957)) of S. anglica is changing more rapidly than in these, and other older, species. The die-back of swards in the lower zones of marshes in the south of

A CENTURY OF EVOLUTION IN SPARTINA ANGLICA

76-

- 50

0

-z 4E

!

Vertical range of Spurtinu

I

, Extreme lower limit of Spurlina '(a few isolated plants present)

tt

Mean high water springs (MHWS) A

Mean high water neaps (MHWN )

P

I

0

I

3-

I

2-

1-

+I I

1

I

2

4

6

I

8

I

I

10

12

Spring tidal range ( m )

Fig. 10. Vertical range of S. anglica marsh and spring tidal range for 32 sites from 20 estuaries. (From left to right: Poole Harbour (two lines. the first from Ranwell er al., 1964), Foryd Bay. Teign, Dovey. Mawddach. Traeth Bach. Tamar, Tavy, Milford Haven (two), Tywi, Taf, Red Wharf Bay, Conwy, Lavan Sands, Loughor, Dee, Ribble, Mersey, Shepperdine, Morecambe Bay, Severn (ten sites)).

England and elsewhere, and the changes resulting from competitive interactions at the upper limit throughout its range, are part of this highly dynamic process as the niche of the species evolves.

D. Interactions with Other Species 1. Competition with Puccinellia maritima As described above, the invasion of established S. anglica swards by other vascular plants has been observed throughout the species' range. The major perennial occupant of the lowest salt marsh zones in western Europe prior to the arrival of S. anglica was P. maritima (Adam, 1981; Dijkema, 1984). In some areas S. anglica has invaded former P. maritima-dominated vegetation, and is perceived as a threat to salt marsh pasture (Ranwell, 1972), whilst in others P. maritima invades and

40

A . 1. G R A Y E T A L .

replaces S. anglica. The tendency for P. maritima to replace S. anglica in more northern marshes in Britain is also seen in the Netherlands, where there is a marked contrast between the Waddensea marshes in the north and those of the Delta area in the south west (Beeftink, 1977; Dijkema, 1984; Scholten and Rozema, 1990). The north/south variation in both Britain and the Netherlands is unfortunately largely confounded with variation in soil type and grazing pressure. The northern marshes are generally sandier and more heavily grazed. Recently, Scholten and Rozema (1990, and Scholten et al., 1987) have examined competition between S. anglica and P. rnaritima in the field and laboratory. Removal experiments in the area of the species’ overlap on a Waddensea salt marsh indicated that S. anglica growth is suppressed by P. maritima at the higher levels (showing a significantly higher biomass production when P. maritima was removed), and vice versa lower down. The interacting effects of soil type (sand vs clay), salinity (saline, 400mM NaCl vs brackish, 200mM NaCI), and soil moisture (dry vs waterlogged) were analysed in a competition experiment using a replacement series design (de Wit, 1960) and the effects of nutrients and grazing were investigated by adding KN03 and KH2P04 and by clipping. The overall results indicated that P. maritima was a better competitor under most conditions, gaining an advantage by early growth that pre-empts space and shades emergent S. anglica shoots. The advantage of P. maritima increased progressively with time, especially on sand. Spartina anglica grew better on clay than on sand, and growth on clay was better under dry conditions, whereas on sand it was better under waterlogged conditions. Saline conditions reduced growth in both species more or less equally, added nutrients had little effect on either monocultures or mixtures, and clipping almost completely suppressed the growth of young S. anglica shoots, whereas P. maritima was less affected, especially on brackish dry sand (Scholten and Rozema, 1990). These experiments accord well with field distributions and help to explain how P. maritima is able to invade an S. anglica sward, particularly under lower spring temperatures and on sandier soils. S. anglica also facilitates the establishment of P. maritima (and other species) by the protection it provides against uprooting by tidal currents, the increased surface elevation by accretion of silt around the shoots, structural improvement of the upper layers of sediment by litter accumulation, and by radial oxygen loss from the rhizomes (Rozema et al., 1985; Scholten and Rozema, 1990).

2. The Ergot Fungus Claviceps purpurea In recent years, and especially during the past decade, many populations of S. anglica become heavily infected with Claviceps purpurea (Fr.)

A CENTURY OF EVOLUTION

41

IN SPARTINA ANGLICA

Tul., a pyrenomycete fungus which causes ergot disease. The fungus is most visible during the sclerotial phase when the sclerotia, or ergots, may be seen protruding from the inflorescence (Fig. 11). Overwintering on or below the mud surface, the ergots germinate in late spring and, via a primary infection by the sexual ascospores, the fungus can spread rapidly through the summer by means of asexual conidia extruded from infected florets in a sticky honeydew. Claviceps displays a certain amount of host specificity (Mantle, 1980), and Loveless (1971) suggested that the strain parasitizing S . anglica probably differs sufficiently to be regarded as a distinct taxon. Spartina anglica ergots have exceptionally high alkaloid content (0.91%), consisting of the lysergic acid derivatives ergotamine and ergotoxine (Mantle, 1969). Although reported on S. alterniflora in the USA as long ago as 1895 (Eleuterius, 1970), C . purpurea does not appear to have been recorded in Europe before 1960, where it occurred at a low level on S. x townsendii near Dublin (Boyle, 1976b). Several records of light infections in the early 1960s are recorded before the first heavy infections, in Poole Harbour in 1971 and in Ireland in 1975 and 1976 (Boyle, 1976b;

k

Germinating sclerotium (May/June)

9-10 months Overwintering

II

P.

3 A

__I__

9 - 0 WBBKS

and pre-chilling

(June/July)

L

Sexual oscospores

Differentiation

Primary infection bv

jf

7-10 days

4

(July - November 7)

Sphacelial stage extrusion of honeydew containingasexual conidia

Fig. 11. Life cycle of C. purpurea on S . anglica (reproduced with permission from Gray et al., 1990b).

42

A . .I. G R A Y E T A L .

Gray et al., 1990b). By the 1980s, heavy infection was apparent in several areas including Poole Harbour, the Dee and the Ribble estuaries. In 20 “populations” (each comprising a 0.5 ha area of S. anglica sward) in Poole Harbour, infection rose from a mean of 36.4% of all inflorescences containing at least one ergot (or honeydew) in 1983 to a mean of 85.2% in 1988 (Gray et al., 1990b). More than 90% of all flowering heads were infected in several populations. In the Ribble estuary, infection rose from less than 1% of infected inflorescences in 1981 to as high as 52% in the main sward in 1986, and similar increases were recorded in the Dee between 1986 and 1988 (Thompson in Gray et al., 1990b). These exceptionally high, even epidemic, levels of disease are relatively rare in natural plant communities (Burdon, 1987; Clarke e[ a l . , 1987) and prompt at least two important questions: what are the effects of the pathogenon the host? and why are infection levels so high? The first question is difficult to resolve. Although there is a clear effect on fecundity, in that every ergotized floret reduces potential seed production by one, the extremely low levels of seed-set in most populations and the possible effects on seed production in uninfected florets make measurement of the actual effect on fecundity extremely difficult. An estimate for Poole Harbour populations in 1985 indicated that potential seed production was reduced by 16% on average, but by as much as 47% in some populations, and there were eight ergots for every filled spikelet (Gray et a l . . 1990b). In addition, there is limited evidence that the demand of the fungus on the host’s resources may lead to host damage-a significant negative relationship between mean ergot weight and number of ergots per inflorescence suggesting that developing sclerotia compete for limited nutrients. The extent of the effect is impossible to measure in field populations because it is not possible to recognize individual plants (ramets). The uncharacteristically heavy infection in S . anglica is most obviously related to the fact that the species tends to occur as dense, monospecific swards, and thus resembles an agricultural crop in its vulnerability to epidemics (Burdon and Chilvers, 1982). Furthermore, the apparent genetic uniformity of S. anglica, its proliferation mainly by clonal expansion, and its extended period of flowering from July to November make it especially vulnerable to pathogen attack. Speculation as the the future development of the disease must take account of the species’ current inability to evolve resistance, of the possible effects of unusual weather conditions which might break the cycle of infection, and of the recently discovered presence of a hyperparasite, the fungus Fusarium hererosporurn, which is a potential agent of biological control of Claviceps (Gray et a l . , 1990b).

A CENTURY OF EVOLUTION IN SPARTINA A N C L I C A

43

3. Invertebrates The fauna of S. anglica marshes has received very little attention, in contrast to that of the S. alterniflora-dominated marshes of North America (e.g. Teal, 1962; Daiber, 1982). Payne (1973) noted only four insect species commonly occurring in S. anglica swards in Poole Harbour. These were a leafhopper, Euscelis obsolerus, a herbivorous grasshopper, Chorrhippus albomarginatus, an omnivorous grasshopper, Conocephalus dorsalis, and a predatory damsel bug, Dolichonabis linearus. The first three appeared to be feeding on S. anglica, although C. dorsalis fed also on E. obsolerus, as did D. lineatus. This rather skeletal insect food-chain mirrors, but is less rich than, S. alterniflora marshes. Other insects found in Poole Harbour S. anglica marshes included three species of shorebug (Saldidae) and several spider species, including Argiope bruennichi. Jackson and his colleagues (Jackson, 1984; Jackson et al., 1985, 1986) also found relatively few species of macroinvertebrates in the S. anglica canopy of a marsh at Seafield Bay in East Anglia-13 taxa in all, and only six occurring regularly. Of these, only one, the sap-feeding spittlebug Philaenus spumarius, was a significant consumer of live S. anglica material, and the consumption of this species was a very small fraction of the plant's annual production. The total annual assimilation by all canopy invertebrates amounted to less than 0.3% of the total annual above-ground net primary production, with P. spumarius assimilating 0.4g of a canopy-species total of OSgcm-*year-'). As in North American marshes, the assimilation of Sparrina material was dominated by sediment-associated detritivores (assimilating more than 40 gcm-' year-'), of which the dominant species was the polychaete worm Nereis diversicolor, accounting for more than 85% of all S. anglica material assimilated. Although dominated quantitatively by N . diversicolor, the invertebrate fauna associated with the sediments in this marsh was surprisingly rich in species, at least in the lower zones, with a total of 15 species of which 12 occurred regularly. Although this is a similar level of diversity to S. alterniflora marshes, Jackson et al. (1985) suggest that the inequitable density distributions, in which one species dominated, reflects the recent origin of S. anglica marshes. Observations elsewhere (Millard and Evans, 1984; S. McGrorty, pers. comm.) indicate that the benthic fauna of advancing or stable S. anglica marshes is generally depleted in relation to the nearby tidal flats, and that although epibenthic species may be present in S. anglica swards, the Seafield Bay site may not be characteristic. Interestingly, because it is the only C, species present, Jackson's study of the fate of S. anglica in the food chain was able to exploit the

44

A . J . GRAY E T A L .

differences in carbon isotope utilization between C3 and C4 species and the tendency for animals to reflect the I3C/l2C ratio of their food sources (Teeri and Schoeller, 1979). His finding that few of the canopy species utilize S. anglica as a major food source may reflect the relative indigestibility of C4 species as a whole (Caswell and Reed, 1976) or the relatively recent origin of the species. Of the annual net primary production of S. anglica at Seafield Bay, between 14% and 20% was assimilated by invertebrates (> 85% of that by N . diversicolor), mostly as detritus in situ, about 30% was dissipated by micro-organisms and meiofauna, and of the rest around 15-22% is exported away from the site as detritus or live material and is available for incorporation into estuarine food-chains (Jackson, 1984).

4. Spartina anglica and the Decline of the Dunlin. Calidris alpina Whatever its richness and density, the invertebrate fauna of S. anglica swards is much less available as food to wading birds (Charadrii) than that of the nearby tidal flats. Indeed, the major threat which S. anglica is perceived to pose in European estuaries is its invasion of bird feeding-grounds-either the eelgrass (Zostera spp.) and algal (Enteromorpha spp.) beds utilized by herbivorous wildfowl such as the Brent goose, Branta bernicla, and widgeon, Anus penelope, or the invertebrate-rich mudflats used by waders (Doody, 1990; Nairn, 1986; Ranwell and Downing, 1959). Despite this perception, documented evidence of an effect on birds is rare, and the overwintering populations of most species of waders have remained constant or have increased in size since nationwide counts began in the early-1970s (Goss-Custard and Moser, 1988). A notable exception is the dunlin, Calidris alpina, the overwintering numbers of which had declined up to 1988 by almost a half since 1973/74 (Salmon and Moser, 1985). The dunlin was among the species which fed particularly at the tide edge and on the higher parts of the flats in the Dyfi estuary and whose numbers declined during the period of S . anglica expansion there in the early and mid-1970s (Davis and Moss, 1984). (The other species were oystercatcher, Haematopus ostralegus, ringed plover, Charadrius hiaticula, and sanderling, Calidris alba .) Millard and Evans (1984) also noted that dunlin, and other flock-feeding species, prefer to feed on open mud and, unlike redshank, Tringa totanus, were rarely found among S . anglica clumps or in the sward at Lindisfarne in northern England. Goss-Custard and Moser (1988, 1990) compared the different rates of decline of dunlin numbers in different British estuaries to the changes in abundance of S. anglica in those estuaries during the period 1971/72 to 1985/86. Numbers had declined at the greatest rate in

A CENTURY OF EVOLUTION IN S P A R T I N A A N G L I C A

45

estuaries where S. anglica had expanded most during that period and, with some exception, had hardly changed in those where S. anglica populations were static. It is suggested that S. anglica both removes feeding areas and reduces feeding time, both of which would increase rates of emigration and mortality. However, as the authors point out, the decline of dunlin and the spread of S. anglica may have occurred independently, or may be linked by a third factor such as sedimentary or nutritional changes in the estuaries. The removal of S. anglica using herbicides has resulted in the return of waders, including dunlin, to formerly vegetated areas at Lindisfarne (Corkhill, 1984; Evans, 1986), but areas from which the grass had died back along the English south coast are not being exploited. This may be because these sediments are unsuitable for invertebrates to colonize or are covered with mats of algae (Tubbs, 1977).

E. Die-back, Control and Conservation Although S. anglica spread rapidly in the early decades of this century and continues to spread in western and northern estuaries, the degeneration of swards along the south coast of England was noted as early as the mid-1920s. Patchy degeneration was reported from the Beaulieu estuary in 1928 and by the mid-1950s about 60 ha of S. anglica marsh had been lost in the Hampshire Basin (Manners, 1975). Die-back began in Poole Harbour in the mid-l920s, where S. anglica was first recorded in 1890, had spread to cover more than 775 ha by 1924. but had declined to around 415 ha by 1980 (Gray and Pearson, 1984). Other detailed studies of decline have been made from Langstone Harbour (Haynes and Coulson, 1982) and from Milford Haven in South Wales (Dalby, 1970; Baker, 1976; Baker et a f . , 1990). Decline in area of S. anglica swards has been recorded from many English south and east coast estuaries and from some in south west Britain, from northern France and from the south west Netherlands. Investigations into die-back, mainly carried out at Southampton University between 1953 and 1965, did not conclusively provide a proximal causative agent but clearly characterized the conditions under which die-back occurs (Goodman, 1960; Goodman and Williams, 1961; Goodman et a f . , 1959; Sivaneson and Manners, 1972). These include badly drained, usually waterlogged, highly anaerobic soils with a high proportion of fine particles and a high sulphide content. The high organic levels which result from initial S. anglica decay may increase the rate of die-back by increasing the water-retaining ability of the muds. Both toxic levels of sulphide (Goodman and Williams, 1961) and anoxia

46

A . J . GRAY ET A L .

in the rhizomes (Barker, 1963) have been directly implicated in the death of plants in die-back areas. Whatever the immediate cause of die-back, its interest from a broader perspective comes from the series of physiographical and physical changes which follow the invasion of S. anglica, particularly the creation by rapid sediment accretion of a series of badly drained marshes, often with a concave profile. Replacing more gradually sloping mudflats, these marshes remove from circulation extremely large volumes of sediment. Depths of up to 1.8 m of accreted sediment have been recorded below S . anglica marshes in Poole Harbour (Hubbard and Stebbings, 1968). Thus, when die-back begins, the sediments (and presumably nutrients) released are likely to have a major impact on the intertidal area to seaward. This is illustrated by the changes in the bed levels of major navigating channels within Poole Harbour between 1849 and 1980, during which period S . anglica invaded and subsequently died back (Fig. 12). Die-back can therefore be viewed as a ‘natural’ process in which a newly evolved species has dramatically altered the sedimentary and

-800

c m

-600

v

-

.-

L

m

P

-

II]

0

-400

m U

-200

-0

Fig. 12. The percentage change in bed levels of four major channels in Poole

Harbour. Dorset, between the Admiralty Surveys of 1849 and those of 1934, 1954 and 1980. (-m- Wytch Lake, -0- Middlebere Lake, -0- Wareham Channel, -0- South Deep). Also shown is the change in area of S. anglica marsh (. . . A . . .). Modified from Gray et a/., 1990a.

A CENTURY

O F EVOLUTION IN S P A R T I N A A N G L I C A

47

drainage characteristics of the marshes, and created the anaerobic, waterlogged conditions which led to its own destruction. Whether this process will occur on the generally sandier marshes in the west and north, and whether it will be a cyclic process in the south, S. anglica reinvading newly accreted mudflats, are extremely interesting questions (the lack of reinvasion thus far may be because mudflat levels are too low or unstable). The answers to them will determine the eventual niche which the species occupies. Whilst extensive die-back occurs in the south (and several local agencies responsible for sea defence contemplate replanting S. anglica to protect sea walls), elsewhere the continued spread of the plant has prompted attempts to control it. Of ten control programmes reviewed by Way (1987), four were to prevent encroachment onto amenitybeaches, three were to prevent colonization of bird-feeding areas, two were to preserve the floristic diversity of adjacent salt marsh, and one was to clear a harbour prior to the creation of an amenity beach. Most programmes involved a combination of physical removal (by digging or bulldozing tussocks and clumps) and herbicide spraying, either with a backpack, a tractor-mounted boom, or aerial spraying. The most commonly used herbicide has been glyphosate (Roundup@), and the most frequently recommended has been Dalapon@,a sodium salt of dichloropropionic acid. Way (1987) suggests there is a need to conduct trials using low dosages of more modern herbicides. Only partial success in controlling the plant is reported from most areas, and eradication once it has become established is extremely difficult and costly. However, providing a long-term commitment is made, S. anglica can usually be contained by repeated spraying. Significantly, half of the reasons given above for controlling S. anglica relate to aspects of nature conservation. In deciding whether the species was harmful or not to nature conservation interests. Doody (1990) lists its beneficial effects as: 1. 2. 3. 4. 5.

preventing coastal erosion and stabilizing mudflats; aiding reclamation for agriculture: its high productivity in the estuarine ecosystem; the creation, via succession, of grazing marsh; and its value for research.

Against this, he balances the harmful effects of 1 . invading intertidal flats rich in invertebrates and utilized by overwintering waders and wildfowl; 2. replacing more diverse plant communities; 3. producing dense, monspecific swards which alter succession and

48

A. J. GRAY E T A L .

are replaced in ungrazed areas by equally species-poor communities: and 4. promoting agricultural reclamation, and thus the destruction of species-rich, high-level salt marsh. On balance, Doody concludes that S . anglica must be regarded as a threat in estuaries of high wildlife interest, both to bird populations and to ‘natural’ salt marsh succession.

VII. THE FUTURE OF SPARTZNA ANGLZCA Human intervention has played a key role in the initial speciation process which created S. anglica and in the subsequent success of the species through extensive planting. It is also clear that the intervention of man, both direct and indirect, is likely to have a major effect on the immediate future of S . anglica and on its long-term evolutionary survival. Human intervention apart, it is likely that the two main factors that will determine the long-term success of the species are its occupation of a vacant niche which exists on the seaward edge of salt marshes and its high levels of phenotypic plasticity (Thompson, 1989). It appears likely that the species will continue to expand its total area in Britain and northern Europe to occupy much of the potential salt marsh habitat that is available to it, subject only to the continued loss of salt marsh resulting from estuarine development or rising sea levels. On the basis of current evidence, attempts to limit this spread on a large scale, or even to totally eradicate S. anglica by extensive treatment with herbicides, are futile. Once the species is established there appears to be little that can be done to eradicate it totally from an estuary, and the development of a management plan for salt marsh habitats and estuaries which takes account of its presence would appear to be the only economically sensible strategy. The full extent of the possible range that is available for colonization by S. anglica remains to be determined. The success of introductions in northern Europe and as far afield as Australia (Ranwell, 1967) and China (Chung, 1990) suggests that the species has the potential to occupy and dominate the equivalent niche on mudflats in estuarine environments throughout the cool temperate regions of the world, subject only to competition from other Spartina species and climatic limitations. Though the physiological evidence strongly suggests that there will be a climatically defined limit to its northern expansion through the influence of cool spring temperatures which limit early season growth (Long, 1983), there is as yet no clear information available with which to predict the extent of its southern

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expansion into warmer climates, although Ranwell (1967) does suggest that winter temperatures may be important. The rate at which future expansion of S. anglica may take place is difficult to predict, Whilst direct human intervention has been the major factor in its long-distance dispersal within the short timescale over which it has existed, there is clear evidence that, once established, the species is able to spread efficiently and occupy an individual estuary (although there may be a significant lag phase from its original colonization to its full occupation of all the available sites). Because of the widespread planting of the species throughout the British Isles. it is not yet apparent how efficient it is at independent dispersal between estuaries through seed or clonal fragments unless the estuaries are adjacent or are linked by intermediate areas of salt marsh. However, it did spread successfully, apparently unaided, more than 150 km eastwards along the coast and more than 100 km across the English Channel from its point of origin in the early part of the century. The efficiency of such “independent dispersal” will be important in determining the rate at which the species can expand to colonize available habitats and will be a major factor in the future spread of the species from introduced sites throughout the world. In the longer term it appears likely that the process of global warming may have a major effect on the future of S. anglica. There are two main consequences of global warming which may have a direct impact. The predicted rise in sea levels will result in the destruction of some coastal habitats through inundation, and although new ones will replace them, a long time may pass before new stable habitats become established (Rowntree, 1990). The second likely consequence of global warming is that S. anglica (in common with many other species, e.g. see Grime (1990)) may alter its geographical range and become of increasing importance in north European estuaries as temperatures increase and allow improved early season growth at northern latitudes. The likely timescale of such predicted changes in global temperature and erosion of salt marsh in the face of rising sea level will be important in determining how well S. anglica is able to repeatedly recolonize available habitats in the face of such dynamic changes to the environment. The success of S. anglica so far is due in part to its great flexibility, its highly plastic phenotype enabling it to tolerate a wide range of substrates and conditions, including high accretion rates. It is at least tempting to speculate that this flexibility may in part be due to the high levels of fixed heterozygosity which exist in such a high polyploid of hybrid origin. Although such flexibility may account for its short-term success, the evolutionary success of the species on a much longer timescale will

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depend greatly on its ability to generate genetic variation. Such variations will be required to enable it to evolve adaptively in response to evolutionary pressures from pests and pathogens. Since all the current evidence suggests that the existing levels of variation are extremely low as a direct consequence of the genetic bottleneck that occurred during the speciation process (Raybould et al., 1990b), S. anglica is faced with a major evolutionary problem. The conventional processes, such as mutation, which create suitable adaptive variation for a species, occur over a relatively long timescale, though in S. anglica there is some potential for the generation of variation by either the breakdown of preferential pairing or by backcrossing with one or other parental species (see above). S. anglica is nevertheless faced with the evolutionary risk of surviving through a period of severely reduced genetic resources. The current dominance of S. anglica swards by what appears to be a single genotype may be an inherently unstable situation since, like many agricultural crops, it will be vulnerable to epidemics (Burdon and Chilvers 1982). Indeed, it might be argued that the current vigour and dominance of S. anglica in salt marsh habitats may in part be the result of the relative absence of a natural complement of adapted herbivore, pest and disease organisms, as has been suggested, for example, to explain the vigour of Eucalyptus species in alien habitats (Pryor, 1976). This temporary escape from enemies may mean that not only the S. anglica swards themselves but also the entire salt marsh communities which they protect from erosion, may in the long-term be vulnerable to the evolution and spread of Sparrina pests or pathogens. The potential danger of such epidemics is well illustrated by the recent dramatic spread of ergot fungus through many UK S. anglica swards since the early 1960s (see Section VI above). Apart from its immediate effect, this pathogen may have a direct bearing on the evolutionary future of the species by severely limiting seed-set. The resulting reduction in sexual reproduction, though of comparatively minor significance in the shortterm survival of aspecies such as S. anglica, which efficiently reproduces itself asexually, may drastically limit the rate at which the species can evolve to combat successfully this and any future pest and disease epidemics. The future of S. anglica will therefore not only provide us with a unique opportunity to study many important ecological processes as the species and its niche develop together, but also will provide us with an important evolutionary laboratory to investigate the initial processes in the success or failure of a new species. The key elements in the second century of this ecological and evolutionary drama are likely to be the role of human intervention, both direct and indirect, and the ability of

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the species to respond to evolutionary pressures in the face of its current severely limited genetic variability.

ACKNOWLEDGEMENTS We are especially pleased to acknowledge the help we have received from other ‘Sparrina scientists’ including Paulina Benham, Ralph Clarke, Mick Drury, Maggie Hill, Mike Lawrence, Steve Long, Tom McNeilly, Karen Myers, Judith Pearson, the late Derek Ranwell, John Thompson and Liz Warman. We are grateful to Mary Perkins for her hard work in producing the manuscript. The support of the Natural Environment Research Council (grant to A. F. Raybould), the Energy Technology Support Unit of the Department of Energy (funds for work on Spartina niche), and BP Petroleum Development Ltd (student grants) is also much appreciated. Alastair Fitter and an anonymous referee made helpful comments on the text.

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Spartina alterniflora growing in France and Spain and comparison with Spartina x townsendii and Spartina anglica. Doc. Phytosociol. N . S . 2, 273-282. Huskins, C. L. (1930a). Origin of Spartina townsendii. Nature 127, 781. Huskins, C. L. (1930b). The origin of Spartina townsendii. Genetica 12, 531-538. Hussey, A. and Long, S. P. (1982). Seasonal change in weight of above and below-ground vegetation and dead plant material in a salt marsh at Colne Point, Essex. J . Ecol. 70, 757-772. Hutchinson, G. E. (1957). Concluding remarks. Cold Spring Harb. Symp. Quant. Biol. 22, 415-427. Jackson, D. (1984). Salt-marsh populations and the fate of organic matter produced by Spartina anglica. B. E. S . Bull. 15, 192-196. Jackson, D. Mason, C. F. and Long, S. P. (1985). Macro-invertebrate populations and production on a salt-marsh in east England dominated by Spartina anglica. Oecologia (Berl.) 65, 406- 411. Jackson, D., Harkness, D. D., Mason, C. F. and Long, S . P. (1986). Spartina anglica as a carbon source for salt-marsh invertebrates: a study using 13C values. Oikos 46, 163-170. Jones, P. (1980). Leaf mottling of Spartina species caused by a newly recognised virus, spartina mottle virus. Ann. Appl. Biol. 94, 77-81. Jones, M. B., Hannon, G. E . and Coffey, M. D. (1981). C4 photosynthesis in Cyperus longus L., a species occurring in temperate climates. PI. Cell Environ. 4, 161-168. Kamps, L. F. (1962). Mud distribution and land reclamation in the eastern Wadden shallows. Rijkswaterstaat Commun. 4, 1-73. Konig, D. (1948). Spartina townsendii an der Westkuste von Schleswig-Holstein. Planta 46, 34-70. Lambert, J. M. (1964). The Spartina story. Nature 204, 1136-1138. Lawrence, M. J. (1984). The genetic analysis of ecological traits. In: Evolutionary Ecology (Ed. by B. Shorrocks), pp. 27-64. Blackwell, Oxford. Lee, W. G. and Partridge, T. R. (1983). Rates of spread of Spartina anclica and sediment accretion in the New River estuary, Invercargill, New Zealand. N . Z . J . Bot. 21, 231-236. Long, S. P. (1983). C4 photosynthesis at low temperatures. PI. Cell Environ. 6 , 345 -363. Long, S . P. (1990). The primary productivity of Puccinellia maritima and Spartina anglica: a simple predictive model of response to climatic change. In: Expected Effects of Climatic Change on Marine Coastal Ecosystems (Ed. by J. J. Beukema, W. J. Wolff and J. J. W. M. Brouns). pp. 33-39. Kluwer, Dordrecht. Long, S. P. and Incoll, L. D. (1979). The prediction and measurement of photosynthetic rate of Spartina townsendii in the field. J . Appl. Ecol. 16, 879-891. Long, S . P. and Woolhouse, H. W. (1978). The responses of net photosynthesis to light and temperature in Spartina townsendii (senw lato), a C, species from a cool temperate climate. J . Exp. Bot. 29, 803-814. Long, S. P . , Incoll, L. D. and Woolhouse, H. W. (1975). C4 photosynthesis in plants from cool temperate regions with particular reference to Spartinn townsendii. Nature. 257, 622-624. Long, S. P., Dunn, R., Jackson, D., Othman, S. B. and Yaakub, M. H. (1990). The primary productivity of Spartina anglica on an East Anglian estuary. In:

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Genetic and Phenotypic Aspects of Life-history Evolution in Animals R . H . SMITH

...................................... .................................... A . The fieris rupae Problem ....................... B . Aims of This Review . . . . . . . . . . . . . . . . . . . . . . . . . 111. Conceptual Frameworks . . . . . . . . . . . . . . . . . . . . . . . . . . . . A . Life-history Theory ........................... B . Quantitative Genetics ......................... C . Multivariate Selection ......................... D . Gene-Environment Interactions . . . . . . . . . . . . . . . . . . E . The G and E Matrices ......................... IV . Experimental Approaches Illustrated by Cullosobruchus . . . . . . . A . Phenotypic Correlation ........................ B . Experimental Manipulation ..................... C . Breeding Designs ............................ D . Selection Experiments ......................... E . Mapping the Options Set ....................... V . Discussion., . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I . Summary

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I SUMMARY Fitness is determined by survival rate and fecundity at different ages . Life-history theory is concerned with how constraints lead to optimal combinations of age-specific survival and fecundity. that is combinations which maximize fitness in a given environment. The optimization approach emphasizes constraints of resource allocation. often described as physiological trade-offs . Experimental examination of trade-offs has 0

ADVANCES IN ECOLOGICAL RESEARCH VOL 21 ISBN 0-12-013921-9

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often involved physiological manipulations. In contrast, geneticists emphasize genetical constraints (genetic correlations resulting from pleiotropy and linkage, and epistatic interactions between loci). Genetical constraints are conventionally examined by quantitative genetic analysis. Some recent papers have successfully reconciled the two approaches, and in particular Charlesworth (1990) has explicitly examined the relationships between functional constraints on life-history characters and the genetic variances and covariances at equilibrium. In this paper the optimization approach is exemplified by the general iteroparous life-cycle model of Sibly and Calow (1983). The quantitative genetics description of life-history evolution is developed in relation to a bivariate case of the general iteroparous life-cycle model, and the recent results of Pease and Bull (1988) and Charlesworth (1990) are outlined. The importance of gene-environment interactions is stressed, especially in relation to variation in population density. Although the theory of negative pleiotropy as a means of maintaining genetic variation in components of fitness has become widely accepted, it is noted that positive genetic correlations between constrained variables at equilibrium can arise in the multivariate case, depending on the pattern of relationships between functional constraints. Different experimental approaches to investigating life-history evolution in animals are compared by reference to a series of papers on the seed beetle Callosohruclius macirlatus, a semelparous species with a relatively simple life-history. The data illustrate some of the predictions of multivariate selection theory, including the existence of both positive and negative correlations between components of fitness at genetic equilibrium. Also, it is shown that experimental manipulation can reveal a trade-off in resource allocation when genetic analysis failed to reveal a trade-off because of lack of genetic variation. Charlesworth (1990) showed that the multivariate selection model leads to the same predictions of mean phenotype as the optimization approach for both simple optimization and frequency-dependent selection (ESS predictions). It is suggested that frequency-dependent selection and the existence of alternative ESS life histories with different equilibrium genetic constraints may be an important consequence of environmental patchiness.

11. INTRODUCTION Fitness in natural populations is clearly some combination of survivorship to ( I , ) and fecundity at ( m , )different ages. However fitness is defined (Nur, 1987; Caswell, 1989a; Sibly, 1989; Parker and Maynard

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Smith, 1990), maximising fitness by differential selection on combinations of (/,, m,)is the natural tendency of natural selection (Fisher, 1930), even though the direction of selection may change because of temporal changes in the external environment, population density (Charlesworth, 1980) and the frequency of different genes (Maynard Smith, 1989a,b). One of the guiding principles of theories of life-history evolution is that you cannot get something for nothing, that is an increase in one component of fitness is expected to be at the expense of another (Sibly and Calow, 1986). Geneticists stress genetic constraints (Loeschke, 1987) that act on top of physiological or other functional constraints (Maynard Smith, 1989b), and one important issue is the relationship between genetic and functional constraints (Charlesworth, 1990). Another concerns genetic variation in fitness (supposed to disappear inevitably under natural selection; Fisher, 1930) as opposed to variation in components of fitness (Smith et a l . , 1987). In this area, ecologists have gained a great deal from the accumulated theory and experience of animal and plant breeders (e.g. Falconer, 1989a,b). In practical breeding, a selection index is often established, based on genetic constraints (genetic correlations) as well as an economics-linked quantity to be maximized; in life-history evolution, the quantity to be maximized happens to be fitness. Therefore artificial selection in the context of, for example, animal breeding has close analogies with life-history evolution (Hill and Mackay, 1989).

A. The Pieris rupue Problem Evolutionary biologists tend to accept that life is governed by neo-Darwinian principles, perhaps rather uncritically (Gould and Lewontin, 1979). Few biologists question the central dogma of neo-Darwinism, that evolution is the process and consequence of maximizing individual fitness. However, in a series of papers on the small white butterfly Pieris rapae L., Gilbert (1984a,b,c, 1986) suggested that his data on the distribution of size and fecundity in field populations of P. rapae were inconsistent with conventional evolutionary theory. The arguments can be summarized as follows:

I . Gilbert's Thesis ( a ) Observation: body size is heritable. In a population of butterflies recently derived from wild-caught parents, there was a significant regression of pupal weight of offspring on pupal weight of parents. From the slope of the regression, the heritability of pupal weight was estimated as h 2 = 0.45 (Gilbert, 1984a). Heritability h' measures the proportion of phenotypic variation in a character expressed in the test

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environment which results from genetic differences between individuals (Section 1II.B). Thus there must have been considerable genetic variation between individual butterflies to give rise to the observed heritable variation in body weight. Gilbert (1984a) confirmed that the genetic variation in body weight was additive (Falconer, 1989a) by selection experiments on his base-line strain (B). Selecting upwards for large pupae produced a strain (C) in which mean pupal weight was 20% higher, and selecting downwards produced a small strain (A).

(b) Observation: body size is related to fecundity. Pieris rapae follows the general rule in insects that heavier females lay more eggs (e.g. Credland et al., 1986; Li, 1988; Smith and Lessells, 1985; Wrelton, 1987). In the unselected line (B), Gilbert (1984a) found a correlation between pupal weight and lifetime fecundity and predicted that butterflies in the large strain (C) had a potential average increase in fecundity of 20%. Note that this prediction is based on a simple (phenotypic) correlation between fecundity and weight.

(c) Deduction: there is heritable variation in a component of fitness. Putting together observations 1 and 2, Gilbert (1984a) concluded that there was heritable variation in fecundity, which is a component of fitness. Gilbert found no evidence of a trade-off in other components of fitness, and concluded that his results contradicted Fisher’s (1930) Fundamental Theorem of Natural Selection which says that the rate of evolution is proportional to additive genetic variance in fitness, with the corollary that a population that is not evolving should have no genetic variation in fitness.

2 . The Optimality View Smith et al., (1987) suggested an alternative interpretation of the P. rapae data. Their approach was first to derive an expression for fitness that included the major events in the life-cycle of P . rupae, then to establish what quantitative relationships would be necessary to balance (or trade-off) components of fitness, and finally to examine the data to see whether there was evidence of such trade-offs. In their re-analysis of Gilbert’s data, they found some evidence for a trade-off between fecundity and juvenile survival (the selected, large C line had a higher juvenile mortality rate; see their (Fig. l ) , but the results were not consistent for both selection regimes (the downwards-selected, small A line also had a higher juvenile mortality rate). In their discussion, they suggested how a genetic trade-off might be identified by searching for a negative genetic correlation (the negative pleiotropy of Rose, 1982), but no data were available to do this.

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3. Comment In fact, neither Gilbert (1984a,b,c) nor Smith et ul. (1987) had established that there was a problem about fitness to address, since the correlation between weight (which was heritable) and fecundity (which may or may not have been heritable) was phenotypic; no genetic correlation was established then or since (Moller et al., 1989a), though the relationship between genetic and phenotypic correlations highlighted by Bell and Koufopanou (1986) suggests that there was likely to be a positive genetic correlation (see Section 111) and therefore that there probably was also heritable variation in fecundity. The debate about P. r a p e highlights some important questions which will be discussed below. These include: (i) can optimality theory help us to understand life-history evolution? (ii) what is the relationship between fitness and its components? (iii) how do genetic constraints relate to the underlying trade-off constraints assumed in optimality theory? (iv) how will evolution proceed under genetic constraints, and does it matter whether all the constrained variables have been measured? (v) what data must be recorded to test the theory?

B. Aims of This Review There is now an extensive empirical and theoretical literature relating to the above questions. The main aim of this review is to guide animal ecologists through the important theoretical results and to attempt synthesis where it seems appropriate. The next section sets up the conceptual frameworks, and the generalized iteroparous life-history model discussed by Sibly and Calow (1983, 1986) will be used to link optimality theory with quantitative genetics. The section on experimental approaches is related back to the theory, in particular the recent important paper by Charlesworth (1990). Data collected at the University of Reading in a series of studies (Moller et ul., 1989a,b, 1990; Sibly et ul.,' 1991) on Cullosobruchus muculutus are used to illustrate the utility of different experimental approaches. This case study provides a contrast to the many studies on Drosophilu melunoguster and, being a semelparous species, the fundamental constraints on C. muculutus are more easily understood. The discussion briefly introduces two new topics; frequency-dependent selection and environmental patchiness. These are highlighted as the areas requiring further theoretical and experimental work.

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K. H. SMITH

111. CONCEPTUAL FRAMEWORKS The questions raised by Gilbert (1984a,b,c) relate to two aspects of population biology: life-history theory (questions about the allocation of resources in observed phenotypes) and population genetics (concerned with changes in gene frequency in relation to variation in fitness between individuals). Clearly, both topics are concerned with evolution and yet theory has been developed for both with rather little cross-reference and even less integration until recently (e.g. Roughgarden, 1979). Life-history theorists have mostly used phenotypic optimization models (e.g. Cole, 1954; Charnov and Schaffer, 1973) based on the core concept of the trade-off as a constraint on evolution (e.g. Levins, 1968; Charnov, 1982; Sibly and Calow, 1983). Population geneticists have been concerned with the dynamics of gene frequencies and how they are affected by patterns of genetic correlation, epistasis and pleiotropy. The quantitative genetics extension to continuously varying characters assumes constraints but may not describe them as trade-offs (e.g. Lande, 1979; Falconer, 1989a). There are a few examples in the earlier literature of papers that demonstrate outstanding intuitive understanding of the links between optimization and genetics, for example Lewontin’s (1965) discussion on the effects of selection of life-cycle components on genetic variance in different components of life-history (see also Cole (1954) and Caswell (1978)). Geneticists seem generally to have felt that trade-off constraints are only of evolutionary interest if they are genetically based (i.e. they exist because of negative genetic correlations-Charlesworth, 1984; Reznick, 1985) and that evidence for trade-offs based on phenotypic correlations (e.g. Calow. 1979) is therefore suspect. However, Bell and Koufopanou (1986) have argued persuasively that estimating phenotypic correlations may in practice often be more useful than attempting to estimate negative correlations because there is a range of circumstances where genetic correlations may be inferred directly from phenotypic correlations (Falconer, 1989a. p.315). Mdler et a l . , (1989b) argued that there may be a contingent (phenotypic) response (a physiological trade-off) that complements the evolved (genotypic) response (the genetic tradeoff) yet may not be detected by quantitative genetic analysis; they distinguished between genetic adaptations of, and phenotypic plasticity within genotypes. Pease and Bull (1988) used an explicit linear model of trade-offs to illustrate their discussion of how phenotypic correlations may legitimately identify trade-offs, and Charnov (1989) argued in favour of the optimization approach, and emphasized both the dynamic nature of the genetic constraints (genetic covariances) and that the

LIFE-HISTORY EVOLUTION

69

genetic covariance matrix is a linear approximation of true trade-off relationships that may disguise too much of basic interest (the shape of the trade-off relationship-Bell and Koufopanou, 1986, p.123). The most recent and important paper to relate optimization models to quantitative genetics is by Charlesworth (1990), and will be discussed in Section 1II.C.

A. Life-history Theory

1. Background Caswell (1989a) has reviewed the development of life-history theory since Fisher’s (1930, p. 47) characteristically terse statement of the problem of how much resource to allocate to gonads vs soma. Caswell (1989a) highlighted three important early studies that have stimulated much later research, the first of which is briefly described below. Lack (1947) was interested in why birds lay fewer eggs than they are capable of physiologically, and suggested that there was a constraint on offspring survival because parents are unable adequately to feed and rear too large a number of offspring. Thus there is a trade-off expressed as a negative correlation between realized (as opposed to potential) fecundity and juvenile survival. Lack suggested that the number of surviving offspring would therefore be maximized at an intermediate rather than maximal clutch size, and Lack’s hypothesis has been examined in a number of studies since he elaborated it in his 1954 book. Cody (1966) broadened Lack’s hypothesis to incorporate other calls on an animal’s time and energy. Klomp (1970) reviewed the available data on birds in 1970, and there have been more detailed long-term and manipulative studies of optimal clutch size in a few species since then (e.g. Dhondt et a l . , 1990; Nur, 1984; Pettifor et a l . , 1988). Lack’s hypothesis has been at the core of clutch size models proposed for animals far removed from birds, for example seed beetles (Smith and Lessells, 1985), parasitoids (Charnov and Skinner, 1984; Waage and Godfray, 1985) and general models of insect oviposition (Parker and Courtney, 1984). The models all derive from Lack’s notion of a most productive clutch size (Charnov and Krebs, 1974) which provides a theoretical upper limit to individual clutch size in species that lay more than one clutch of eggs. It should be noted that most tests of predictions relating to Lack’s hypothesis assume that individual animals are able to adjust their behaviour in response to prevailing circumstances (adaptive phenotypic plasticity) rather than being rigidly programmed to lay a certain number of eggs according to genotype, and experimental manipulation has therefore been the main tool used to test the predictions of

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R. H . SMITH

the theory. In one animal (the seed beetle C. maculatus (F.)) there is also evidence of genetic variability in behaviour determining clutch size both between populations (Messina, 1989) and within populations (Messina, 1990); the beetle will be discussed further in Section IV. However, in general, derivatives of Lack’s hypothesis have been tested by examination of phenotypic relationships with varying degrees of rigour (Pease and Bull, 1988). Williams (1957) followed Medawar (1946, 1952) in considering the evolution of senescence in terms of selection acting on individual fitness rather than for the good of the species. Williams’ (1957) paper is described by Charnov (1982) as the type specimen for the use of selection to understand a major biological problem. He blends together physiological and environmental constraints on development and life span, pleiotropy of reproductive function, and a precise definition of soma , . . combined with one simple fitness observation to produce a theory of senescence . . . that reproduction later in life is worth less to an individual simply because the individual is less likely to be alive at the older age.

The idea of an interaction constraint or trade-off between two components of fitness (early reproduction vs later survival) is the primary concept in life-history theory; the importance of timing of life events (Cole, 1954) and the genetic expression of trade-offs as pleiotropy (Rose, 1983a,b) may be regarded as two secondary concepts that are consequences of the first. This broad outline will be developed below in relation to a particular category of optimization model before moving on to other approaches and some case studies in Sections 111 and IV.

2. Trade-offs in Optimization Models The approach used here follows Sibly and Calow (1983, 1986). The primary components of fitness are the probability 1, that an individual survives from age 0 to age x , and the fecundity m,r at age x . It is assumed that there are constraints on the 1, and m, values beyond the trivial constraints 0 s 1, S 1 and m, 3 0. What might be the nature of these further constraints? Clearly there are ultimate constraints determined by the laws of physics and chemistry that are believed to constrain all organisms within the universe and a subset of developmental constraints that prohibit certain combinations of 1, and m, for particular organisms, at least in the short term (Maynard Smith et al., 1985). Because the resources available to an individual are finite, there must be some physiological constraints of the sort envisaged by Williams (1957) and Cody (1966), and these constraints of resource allocation (in part a consequence of constrained resource acquisition) have an obvious

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LIFE-HISTORY EVOLUTION

intuitive appeal as an explanation of observed (or hypothetical) tradeoffs (Gadgil and Bossert, 1970; Calow and Sibly, 1983; Bell and Koufopanou, 1986; Sibly and Calow, 1986). Of course, there may also be various genetic constraints (Gould and Lewontin, 1979; Barker and Thomas, 1987), including the random, temporary loss of genetic variance, but these will not be considered until later sections. Whatever the explanation, the assumption is usually made that an organism is constrained in the values of 1, and rn, that it can achieve such that it only has a limited set of evolutionary options (the options set in Fig. 1). Although the options set (sometimes called the fitness set (Levins, 1968) or trade-off (Charnov, 1982)) has been a key concept in life-history theory, it lacks a precise definition and there is as yet no common agreement on how it should be measured.

3. Defining Fitness Given a set of possible options, some or many of which may actually be observed in a population, which combinations of 1, and m, are expected to emerge as a consequence of natural selection? The combination of I, and m, that is the optimal life history is that which maximizes fitness 1.oo

-N

0.80

00

P

.-G>

0.60

>

5

0.40

0.00 0

2

4

6

8

10

Fecundity (m)

Fig. 1. The options set between two components of fitness, as generally assumed in the optimality approach. The curve that bounds the options set (shaded) is known as the trade-off curve. Fitness is a function of adult survivorship between breeding attempts and fecundity at each breeding attempt, and increases away from the origin.

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R. H . SMITH

subject to the constraints of the options set. Definition of fitness has been a vexed question (Nur, 1987), but there is now a general consensus that the Malthusian parameter or intrinsic rate of increase r defined as the root of the Euler-Lotka equation is the most appropriate (if cumbersome; Parker and Maynard Smith, 1990) measure of fitness (Fisher, 1930; Charlesworth, 1980, 1984; Lande, 1982; Sibly and Calow, 1983; Sibly, 1989); see Sibly and Calow (1986, p. 11) or Maynard Smith (1989a, p. 39) for straightforward justifications. If m,r is taken to be the number of female offspring, the Euler-Lotka equation is:

There are implicit assumptions in the definition of fitness as r (or its discrete-time analogue A = e r , useful for populations structured by stage rather than age; Caswell, 1989b), in particular the demographic assumption of stable age-structure. Caswell (1989b, p. 171) provided an informal justification for the use of stable age-structure theory in the above definition of fitness, as well as a discussion (p. 176) about when maximizing Fisher’s (1930) reproductive value (sometimes presented as a measure of fitness) is equivalent to maximizing r . Up to now, fitness has been defined as a rate of increase without specifying to what that rate refers. Sibly and Calow (1986) specifically defined r as the rate of increase of a dominant gene, in contrast with most other authors who talk about genotypic rates of increase (e.g. Charlesworth, 1984). Population geneticists conventionally formulate models in terms of fitnesses of genotypes, though reformulation using average fitnesses of genes gives equivalent equilibrium results (R.M. Sibly and R.N. Curnow. pers. comm.). In his chapter: The Fundamental Theorem of Natural Selection, Fisher (1930) may have caused some confusion by basing his derivation on the average fitnesses of alleles (e.g. p. 37) but referring to the fitnesses of genotypes in his Summary (p. 50); Crow (1986, p. 84) states very clearly that the genetic variance referred to by Fisher is the variance of the average fitnesses of alleles rather than of genotypes. However, for the purposes of the optimization approach described below, it is sufficient to think initially in terms of comparing phenotypic variants with one another (Sibly and Calow, 1983).

4. Optimal Life-history The approach adopted here is taken from Sibly and Calow (1983, 1986). In a simplified life-cycle, m offspring are produced at each breeding attempt (Fig. 2). Survivorship from when an offspring is produced until first breeding at age t , is s I and survivorship between successive

LIFE-HISTORY EVOLUTION

Time

Survivorship

73

Breeding

Juveniles

t t*

t t*

Fig. 2 The generalized iteroparous life-cycle (after Sibly and Calow, 1986). In the text, t i and t 2 are assumed to be equal, and juvenile survivorship to first breeding is treated as a parameter determined by the environment. Adult survivorship between breeding attempts is assumed to be a decreasing function of fecundity, as in Fig. 1.

breedings at intervals t 2 is s 2 . Then the fitness r is related to the life-cycle variables as follows (Charlesworth, 1980, p. 226). 1 = e-“ls,m + e-“?s2

(2) For simplicity, consider the case where t l = t 2 = 1 (e.g. annual breeders that start breeding at 1 year old) such that: er = slm

+ s2

(3)

Clearly, fitness increases with survivorship of juveniles, survivorship of adults, and fecundity per breeding attempt. There are several assumptions implicit in Eq. (3); two obvious ones that ecologists might object to are that fecundity is independent of age (the rn, curve is horizontal), as is survivorship once offspring are reproductive ( I , is a negative exponential function of age x ) . Equation (3) implicitly assumes what is termed “absorption-costing’’ rather than direct-costing (Sibly and Calow, 1986). Nevertheless, Eq. (3) is a useful metaphor for “the iteroparous life-cycle”, and will be used to make a number of general points in what follows. The life-history characters that define the life-cycle in Eq. (3) can be thought of as elements of a vector z, where z1 = sl, z 2 = s2 and z 3 = rn. The classic “cost of reproduction” trade-off concept is that there is an interdependence between s2 and rn such that if resources are put into increasing m, then less resources are available for s2 which in consequence decreases; the function f l ( z 2 , z 3 )= 0 that relates s2 and m is

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R. H . SMITH

one life-history constraint or trade-off. There could be another constraint, for example f 2 ( z z 3 ) might specify that production of more offspring per breeding attempt leads to a reduction in survivorship of each offspring (a formulation that leads to Lack’s hypothesis about optimal clutch size). Note that in a three variable case there can be no more than two independent constraints else the vector z is fully determined and there is no optimization problem to solve (Charlesworth, 1990). Sibly and Calow (1983) looked at several pair-wise combinations of life-history characters in the more general formulation Eq. (2) in order to predict the progress of life-history evolution in different circumstances. Here, we will take as an example Eq. (3) with only one functional constraint:

(Boundary conditions on s2 require that 1 3 k l s:/(46,).) Juvenile survivorship is taken to be a feature of the external environment, and the constraint f l says that the trade-off curve between m and s2 is convex upwards such that s 2 approaches k , as m approaches 0, and m approaches (kl/bl)lDas s2 approaches 0. Thus f l is the trade-off curve that defines the boundary of the options set for ( s 2 ,m ) . All phenotypes must lie within the options set, defined by the boundary conditions ( m3 0; 0 zz s2 zz 1) and the trade-off constraint ( f l : s2 b l m 2- k l = 0); this options set is shown in Fig. 1. Sibly and Calow (1983) used a simple graphical method to show how the optimal life-history ( s 2*, m*) depends on the juvenile survivorship sl, assumed to be environmentally determined (Fig. 3). From Eq. ( 3 ) , a given fitness r in a given environment characterized by juvenile survivorship s1 is a linear combination of s2 and m . Therefore, lines of equal fitness with slope -sl can be plotted onto the ( s 2 ,m ) plane and the phenotype z* with greatest fitness is the tangent of a fitness contour to the trade-off curve. Figure 3 illustrates the well-known result that low juvenile survivorship should lead to evolution of iteroparity, and high juvenile survivorship to semelparity (Cole, 1954; Bell, 1980; Sibly and Calow , 1983). The graphical solution can be formalized as follows. Maximizing 1 is convenient in this case and is equivalent to maximizing r . The optimal solution must lie on the boundary of the options set, defined by f l . Replace s2 by f l ( m ) = k l - b l m 2 in Eq. (3), then A=e‘ = s l m + f l ( m ) . Setting aA/am = 0 to maximize A gives s1 + ( f l / m )= 0. Hence, for the particular constraint function f l, the optimal life-history is m* = s1/(2b1)and s2* = k , - b,m*2.

+

75

LIFE-HISTORY EVOLUTION

1.oo

0.80

0.60

0.40

0.20

0.00 0

2

4

6

8

10

8

10

Fecundity (m)

1.oo

-(Y

0.80

00

P

5

0.60

b>

.->

0.40 c

3

0.20

0.00

0

2

4

6

Fecundity (m)

Fig. 3 The dependence of optimal life history on environment: (a) low juvenile survivorship (s,) favours low fecundity ( m ) and high adult survivorship ($2); (b) high juvenile survivorship favours high fecundity and low adult survivorship. The parallel straight lines are equal fitness contours, and the optimal life-history (sz*, m*) is starred (after Sibly and Calow, 1983).

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R. H. SMITH

The above graphical method is appropriate for the phenotypic optimization approach, where it is legitimate to consider pair-wise trade-offs between variables taken two at a time. However, it will be seen in Section 1II.C that the genetic formulation requires a more formal description of how natural selection acts on z, the vector of life-history characters. Specifically, it will be necessary to define a selection gradient vector V W , containing the partial derivatives of fitness with respect to each element of the life-history vector z. For Eq. (3), it is easiest first to evaluate the vector of derivatives of A = e' :

VAT = [aA/az]' = [ m ,1, s l ]

(5)

Then since VW = V r = A-'VA (Caswell, 1989a), the elements of V W are: ar/asl = rn/(slm + s2)

ar/as2 = I/(sIrn + s 2 )

(6)

ar/arn = sl/(sIm + s 2 ) It is clear from Eq. ( 6 ) that maximization without trade-off constraints is trivial (sl, s2, m tend to increase indefinitely when a r / a z , = 0 for all z,). Fitness maximization in practice must take into account trade-off constraints (e.g. Eq. (4)). Charlesworth (1990) gave a formal definition of the definition of the optimal solution using the selection gradient VW and the constraints fi(z) on z, and also showed how the optimality criterion (an expression involving the ar/az, and af,/az, terms) is modified by boundary conditions on the z , . At this stage, we simply note that the selection gradient V W specifies the direction of steepest ascent up the surface of increasing fitness r ( t ) and therefore relates to the progress of natural selection (changes in mean phenotype) as well as the long-term outcome (selective equilibrium at z*).

5. Population Dynamics and Life-history Evolution Clearly there is feed-back from life-history evolution to population dynamics via the l,r, m , curves, but population dynamics must also in part determine life-history evolution through definition of the biotic environment. Charlesworth (1980) provided the authoritative account of these links (though few have come to grips with all of Charlesworth's book), and Caswell (1989b) gave an excellent review of more recent results in his treatment of matrix population models. A brief summary of the main features is given below. The life-history characters used in Eq. (3) fit directly into the familiar Leslie matrix model of stage-structured population dynamics: n/+1 =

Ln,

(7)

LIFE-HISTORY EVOLUTION

77

For the iteroparous life cycle described by Eq. (3) there are only two distinct stages, juveniles and adults, and therefore n, and n,+l are two-element vectors containing the numbers of juveniles and adults in a population at times t and t + 1 respectively. The projection matrix L is:

At stable age-structure, the finite rate of increase is the dominant eigenvalue (A1) of L and the stable age structure is the corresponding eigenvector. The second eigenvalue (A,) of L is in tension with A , , and the progress of the oscillatory approach to stable age-structure is more rapid the larger is the difference between A1 and A,. The eigenvalues are the roots of det IL - AZ I = 0, giving

A

= i(S2

f (S:

+ 4mS1)1’2)

(9) Hence we see that the trade-offs that determine the optimal combination of (sl. s2, rn) eventually determine the rate of decline of age-structure oscillations (Caswell, 1989b). Population dynamics may feed-back on life-history evolution by imposing additional, population-level ecological constraints on life-histories. If intraspecific competition constrains population size to be at or close to equilibrium, survivorship and fecundity values are also constrained such that variation in one must be followed by compensatory (density-dependent) variation in another. Sibly and Calow (1986, 1987) refer to this as “ecological Compensation”, and show that some analyses that ignore this complication are not strictly correct for populations at stable equilibrium if the fitness components involved in trade-off constraints show density-dependent compensation. For Eq. (7), stable equilibrium means that ~t,,+~ = n, and the dominant eigenvalue of L is 1. Thus the stable population constraint is ms,+ s2 = 1. Putting msl= 1 - s2 into Eq. (9) gives A1 = 1 and A, = s2 - 1 = -ms,< 0. Hence A2 << Al, and therefore the return to stable age-structure following perturbation will be relatively rapid as a consequence of ecological compensation. The asumption of stable age-structure is implicit in the use of r as a measure of fitness in Eq. (1) (Caswell, 1989b) and it is comforting that there is internal consistency in this part of the formulation for the ecologically realistic restriction of a stable population where A = 1.

B. Quantitative Genetics The essential assumption in all but the most trivial life-history theory is that components of fitness z, cannot evolve independently to their

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selective optima. These constraints on the z , are referred to as trade-offs (Stearns, 1989) and there has been some disagreement on how precisely they should be defined, and how best to measure them. What is the relationship, if any, between apparent trade-offs recorded by observations of phenotypic correlations in unmanipulated populations in the field (Clutton-Brock et a l . , 1983), in the laboratory (Bell, 19d4a,b), and in manipulated populations (Partridge and Farquhar, 1981; Mdler et al., 1989b), or by genetic correlations (Rose and Charlesworth, 1981a,b; Maller el al., 1989a)? Is there a rigorous definition of a genetic trade-off (Smith et al., 1987) as opposed to a physiological trade-off (Calow and Sibly, 1983), and can we say that only one sort of trade-off is worth measuring, or that one sort is always the best (Reznick, 1985; Sibly and Calow, 1986)? Whatever else a trade-off measures, it must involve covariation or correlation in some form. A simple (phenotypic) correlation may be the result of a number of causes, and should ideally be broken down into better defined components. It is always useful to be able to split covariation into different components based on some sort of structuring or classification of data (e.g. intra- and interfamily, or between manipulations). The quantitative genetics formulation is one way of structuring data that leads to definition and estimation of quantities like heritability and genetic correlation that figure so largely in the life-history literature. As shall be seen later, this formulation also throws light on the nature of the non-genetic components of variation and covariation (Section

1II.E). It should be pointed out, however, that quantitative genetics theory is a statistical description of variation that we invoke only because (a) it is useful for some defined purpose, and (b) because we do not know the nature and number of genes underlying what we observe in the phenotype. It is very rare to be able to identify a particular gene or genes that determine variation and covariation in life-history parameters (e.g. Kallman and Borski, 1978). Polygenic models are sometimes presented as if they are some sort of alternative to Mendelian genetics, but of course they are just an elaboration of Mendelian genetics with a number of assumptions about how the different gene loci contribute to the observed phenotype. In fact, the use of polygenic models of inheritance is an admission of ignorance, though it cannot be denied that great strides have been made in animal and plant breeding because of development of quantitative genetics theory (a simple review can be found in Bowman, 1974). In the simplified description that follows, terminology will be based on definitions given by Falconer (1989a). Further details of theory can be found in Kempthorne (1957), Bulmer (1985) and Falconer (1989a).

LIFE-HISTORY EVOLUTION

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1. Assumptions Of the many implicit assumptions, the most important in quantitative genetics are as follows: a character such as fecundity or weight is controlled by the action of many genes at several loci, each with a small, additive effect; the genes are inherited as normal Mendelian genes; there is an environmental effect on a character that is independent of (i.e. uncorrelated with and additive to) genotype; the population is random mating and the relevant genotype frequencies are in Hardy-Weinberg equilibrium; there are no interactive effects either between alleles at a locus (dominance) or between genes at different loci (epistasis). Quantitative genetics theory was developed in the applied fields of plant and animal breeding where these assumptions are not critical provided violations are not too extreme. However, those who believe that a quantitative genetic approach to life-history evolution in some way reveals fundamental truths should beware; the theory is an approximate statistical model that describes phenotypes and should not be treated as anything more.

2. Model A life-history character zi has a value zik in the kth individual in a population with mean Zi described as the sum of total genetic effects gik and environmental effects eik: Zjk

- 21 = g,k

+ ejk

(10)

Hence the phenotypic variance V , in the population is the sum of the genetic variance Vc and the environmental variance V,; Vc and V Eare uncorrelated by definition and therefore there is no covariance term:

v, = v c + v, If there were correlations or an interaction between g, and eik, the effects on V p would as a first approximation be assumed t o be part of V,. Genetic interactions are conventionally separated out of VG (Falconer, 1989a) because only the additive genetic variance V , contributes to long-term evolution (dominance variance V , and epistasis variance V , only affect the process, not the outcome of evolution

v,

= VA

+ v, + V [

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R. H. SMITH

The distinction between additive and interactive effects of genetic variation is important. Fisher’s (1930) Fundamental Theorem of Natural Selection says that the rate of change in mean fitness of a population equals the additive genetic variance for fitness, and some breeding designs used to estimate genetic and environmental components of phenotypic variance (and covariance) do not separate V , from other parts of VG (Section 1V.C). The standard text where these topics are discussed in more detail is Falconer (1989a), while Crow (1986) and Maynard Smith (1989a) also give very good introductions. Evolutionary biologists are often more concerned with correlations of life-history characters with each other and with fitness than with the variance of a single trait. Equations (10) to (12) can be used to describe variation in other traits z j , and the covariance between z , and z j (or between ziand r ) can be described in a similar way:

Caswell (1989a) suggested that the so-called Secondary Theorem of Natural Selection (Robertson, 1968) is a more useful generalization than Fisher’s Fundamental Theorem. The Secondary Theorem relates the rate of change in the mean phenotypic value z, of a trait to the genetic covariance of z, with fitness r :

k ;= CA(Z;,r ) = (a rlaz;)V,( 2;)

(14)

Equation (14) shows why the selection gradient (introduced in Section III.A.4) is so important. Equation (14) will be considered further in Section 1II.C.

3. Dimensions and Dimensionless Quantities: h2 and

rA

Variances and covariances are measured in units that are the square or the product of the units the z,, ti are measured in. This rather obvious point is worth emphasizing since it allows values that some authors seem to treat as pure numbers to be related to the phenotypic means of the populations to which they refer. Charlesworth (1984) used this approach to describe the relative importance of components of genetic variability for several characters in a population of D . melunoguster; he used what might be called a “coefficient of additive genetic variation” ( V A 1 p ( z , ) / 2 ia) ,dimensionless quantity which vaned from 5 to 26% for various characters in Drosophilu in the study described by Charlesworth (1984). The most commonly estimated dimensionless quantity in quantitative

LIFE-HISTORY EVOLUTION

81

genetics is heritability h 2 , defined as the proportion of observed phenotypic variation attributable to genetic differences between individuals in a population. The “narrow-sense” heritability uses only additive genetic variance and equals V A / V p . Published estimates of heritability will not be reviewed here; a recent review by Roff (1990, Table 1.1) gives a reasonably up-to-date list of published heritability estimates for life-cycle and morphological traits in insects. Essentially, heritability estimates measure correlations among relatives to estimate how much of the phenotypic variation measurable in one population at one time is attributable to genetic variation that cannot be measured directly (see Section 1V.C). The value of h2 in practical breeding programmes is to predict a short-term response to artificial selection (Falconer, 1989a). Several authors have stressed the limitations of heritability estimates in understanding evolution by natural selection (e.g. Barker and Thomas, 1987). Keeping V , and V E separate rather than automatically combining them in h2 is always advisable, and the dynamics of life-history evolution depend directly on the variances and covariances (Lande, 1979; Caswell, 1989a; Charlesworth, 1984, 1990) rather than on any ratios derived from them. Here, it is worth noting four important limitations of heritability estimates: differences in h 2 between populations may reflect differences in either V Aor V Eor both; laboratory studies control V E and therefore h 2 estimates may be higher in laboratory studies than in natural populations (Atkinson. 1979); h 2 changes under natural or artificial selection, and therefore predictions based on h 2 are only valid in the short term; if the primary interest is in whether genetic variation is significantly different from zero, V A estimates should be relatively more precise than h2 estimates and therefore more powerful in hypothesis testing; this is because the h2 estimate suffers from sampling variation in both V A and V E . Genetic and environmental correlation coefficients are also dimensionless quantities that are perhaps more useful than h 2 . In particular, the genetic correlation rA is the simplest expression of a genetic constraint on evolution (Clark, 1987) while the environmental correlation r E contains information about resource allocation (physiological trade-offs), though it may be difficult to disentangle from variation in resource acquisition (see Section 1II.E). However, the variance-covariance matrix from which a correlation is derived contains more information than the matrix of genetic correlations and is more directly related to the dynamics of evolution (see Section 1II.C).

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K. H. SMITH

4. Direct and Indirect Selection The primary aim of animal and plant breeding is to use artificial selection to improve some aspect of yield. Direct selection on a trait usually involves choosing a selection of individuals as parents whose mean phenotypic value in a desired character is S units higher than the mean of the population from which they were selected (S is the selection differential). The response to selection R is the difference in mean phenotypic value between offspring of selected parents and the whole of the parental population before selection. Provided there is no additional (e.g. natural) selection on characters correlated with the directly selected character, the response is:

R

=

h2S

(15)

Heritability will change with selection; although it may rise in the early stages, it would eventually fall towards zero if there were no remaining additive genetic variation in the character selected (Falconer, 1989a,b). Directional, artificial selection, however, sometimes leads to no response ( R = 0) even when heritability is significantly non-zero. One explanation relevant here is that the directly selected character may be genetically correlated with components of fitness (fecundity or viability) such that natural selection opposes artificial selection (Falconer, 1989a); that is, there is a genetic constraint on evolution in the selected populations. This explanation is supported by the observation that reversed selection often produces a rapid response even when the response to directional selection has fallen to zero (Falconer, 1989b). Breeders sometimes use genetic correlations to increase the efficiency of selection for a character with low heritability (Falconer, 1989a, p.320). Indirect selection on a correlated character z j with high heritability h ; should yield a more rapid response than direct selection on the character z i of interest if rA > ( h f / h ; ) @ . The relevance of indirect selection and genetic correlations to life-history evolution will be spelled out in Section 1II.C. Falconer (1989a, p.318) gives expressions for the correlated response to selection. Standard formulae involving the selection differential S can be applied to truncation selection (e.g. artificial selection) where only a selected part of a population is allowed to breed. Natural selection, where most or maybe all members of a population breed but with varying success, is less straightforward. However, rearrangement of the formulae (Charlesworth, 1984) gives the correlated response R j of a trait t i to the direct response to selection Ri of a trait z i :

LIFE-HISTORY EVOLUTION

83

Equation (16) can then be used to check that the outcome of natural selection on correlated characters is in accord with estimates of genetic parameters (see Section 1V.D); if there is a substantial discrepancy, then selection may be acting on other genetically correlated characters that had not been considered.

C. Multivariate Selection The concept of a trade-off in optimization theory is clearly related to the genetic correlation in quantitative genetics. When selection favours two traits simultaneously (e.g. two components of fitness), a positive genetic correlation will increase the rate of evolution (the increase in mean fitness) while a negative genetic correlation will slow it down; thus a negative genetic correlation between two traits favoured by selection is sometimes referred to as antagonistic pleiotropy, or negative pleiotropy. Genes that contribute to positive correlations will rapidly become fixed, while those that increase one selected trait but reduce the other tend to be maintained because their “antagonistic” effects cancel out. Thus simple, qualitative theory predicts that negative rather than positive genetic correlations should be found between components of fitness. Some authors have argued that negative pleiotropy is the main or only point of interest in empirical studies of trade-offs (e.g. Reznick, 1985; Rose et a l . , 1987), though we shall see that the multivariate case is more complicated than indicated by the simple bivariate argument above (Pease and Bull, 1988; Charlesworth, 1990).

1. Evolution of Negative Genetic Correlation in Two Dimensions Charnov (1989) used a generalized two-dimensional example to demonstrate the relationship between a functional constraint on two life-history variables and the genetic covariance between them at selective equilibrium, He argued in favour of the optimization approach because, provided there is additive genetic variance in both traits, the population mean values at selective equilibrium are close to the optimization predictions and they should show a negative genetic correlation of - 1 . The terms V A ( z l ) , v A ( z 2 ) and the covariance C A ( z I z, 2 ) must satisfy a particular form because selection pushes up towards the constraint boundary z 2 = f(zl). Using a linearized version of Fisher’s Fundamental Theorem, Charnov (1989) obtained the same dynamical equations for phenotypic evolution as in Lande’s (1982) approach for weak selection. If all the variance V is additive genetic, writing VI1 for VA(zl), V z 2for V A ( z z ) ,V12 for C A ( z lz, 2 ) , the equilibrium condition in the region of selective equilibrium is:

84

R . H . SMITH v22

=

(af/azl)*2vll

=

-(af/aZl)*VII

and v12

where * denotes that a quantity is evaluated at equilibrium. Thus the genetic variance-covariance matrix G at equilibrium has the form:

For the iteroparous life-cycle exemplified by Eq. (3), taking s1 as fixed and considering only s2 and m with the constraint f l , the equilibrium G matrix for z T = [m, s2]is:

This result may seem trivial, but it makes very clear the relationship between the graphical optimality analysis of Sibly and Calow (1986) and the equilibrium genetic analysis. The slope of the fitness contour -sl that determines the optimal solution is the slope of the regression of VA(sZ)on VA(m) at equilibrium. Thus if s 1 is high (predicted to lead to semelparity), we expect the genetic variance of s2 to be larger (relative to the genetic variance in m) than if s 1 were low (iteroparity). However, this result says nothing about absolute values of the genetic variances and covariances at equilibrium, and the equilibrium condition (17) is only satisfied when the curvature of the trade-off curve in the region of the optimal life-history is negligible. The equilibrium conditions are, of course, determined by the selection gradient VW; at equilibrium, the changes Az in the mean values of the z , are zero (Lande, 1979, 1982; Lande and Arnold, 1983) because:

Az

=

GVW

=

GV1A-I

(20)

Since V17 = [ s l ,11 for z T = [m, s2], simple matrix multiplication shows that A z T = [O,O]. What about variance in fitness in this scheme? From Eq. (3): vA(A) = s:VA(m) =

s:VA(m)

+ vA(s2) + 2CA(slmr s2)

+

(21)

vA(s2) -k 2slCA(m, s2)

Substitution from Eq. (19) confirms that there is zero variance in A (and therefore in r ) at equilibrium, Thus the genetic correlation between components of fitness means that variance can be maintained in those components at equilibrium even though the variance in fitness is zero.

LIFE-HISTORY EVOLUTION

85

Charnov (1989) emphasized that the genetic covariance structure is not necessarily fixed and is likely to evolve as selection proceeds; this is not always made clear in other formulations based on Eq. (20) or similar. Also, the G matrix is a linear representation of the underlying trade-off relationship and can therefore give only a limited (and somewhat distorted) view of the underlying trade-off constraint (cf. Bell and Koufopanou, 1986, p. 127). Evolution towards genetic equilibrium is slowed down by its own progress. The negative genetic correlation evolves by differential selection against those genotypes that are not pushed up against the trade-off curve and are not in the region of the optimal life-history (zy, z z ) . As evolution of a negative correlation occurs, the variance in fitness VA(r ) decreases asymptotically towards zero much more rapidly than the variances in fitness components, VI1 and V 2 2 .In a multi-allelic, two-variable, deterministic simulation model based on Eq. (3) and a constraint similar to Eq. (4), L. R. Linton (pers. comm.) found that r, evolved from 0 to -0.9 in a matter of tens of generations, but did not go beyond about -0.95 after several hundred generations. In these circumstances, V , , and V22also decline slowly and the linearized predictions about their relative magnitudes and that of their covariance are approximately correct. However, we should not expect negative genetic correlations always to evolve because of underlying functional constraints, as seems to be implied by proponents of negative pleiotropy (e.g. Rose, 1983a; Reznick, 1985; Rose er al. 1987; Smith et al., 1987; Mgller ef al., 1989a). The two dimensional case is deceptively straightforward, and Pease and Bull (1988) and Charlesworth (1990) have shown that trade-off constraints may combine to lead to positive as well as negative genetic correlations.

2. More than Two Variables The properties of the general multivariate case are too complicated to describe in full here. The theory is a development of Robertson’s (1968) Secondary Theorem of Natural Selection and assumes that z for a given population follows a multivariate normal distribution with a variancecovariance matrix P which is the sum of genetic effects (G) and environmental effects (E). Lande (1979, 1982) is largely responsible for this approach, encapsulated in Eq. (20) and discussed recently by Lande (1988), Turelli (1988) and Caswell (1989b). The problem with generalizing from the two-dimensional case with one constraint is that the relationship between functional constraints and genetic correlations at equilibrium becomes complex. Pease and Bull (1988) noted that a subset of traits comprising a trade-off may have positive covariances even though they are part of a wider genetic trade-off that generates negative covariances with other characters; thus

86

R. H . SMITH

they urged experimentalists to measure too many rather than too few traits in the search for trade-offs. Without writing down specific mathematical functions, it is possible to illustrate this in a rough and ready way (Fig. 4). In both Fig. 4(a) and 4(e), there is the constraint z 1 + z 2 z 3 = k such that all phenotypes lie on or below a plane. However, in Fig. 4(e), an additional constraint of the sort used by Charlesworth (1990) for illustration maintains negative genetic correlations between ( z l , z 2 ) and ( z l r z 3 ) but generates a positive genetic correlation between ( z 2 ,z 3 ) as shown by the projections in Fig. 4(f)-(h)* Charlesworth’s (1990) mathematical formulation is based on n variables with p constraints f, (p < n), so that the components of z can be arbitrarily treated as either constrained or free variables (the n-p free variables are uncorrelated amongst themselves). At equilibrium, the product of the G matrix and the gradient vector is zero (Eq.(20)) implying either that all the variances Vii are zero or that the genetic variance-covariance matrix is singular (there is a linear dependence between the rows and columns of G such that detlGl = 0). Charlesworth (1990) showed that, although some of the genetic covariances V,i must be negative, not all must be, and there are some conditions which imply that some genetic correlations must be positive. Genetic correlations are only constrained to be negative (for constrained pairs of variables) and zero (for pairs of unconstrained variables) if the separate pairs of variables are subject to independent constraints e.g. z I = fl(z2); z 3 = fZ(z4) etc. This shows the limitation of only examining pair-wise trade-offs between variables taken two at a time and highlights the practical problem of deciding how many, and which, variables to measure.

+

3. Selection, Mutation and Non-additive Effects Although the linearized, multivariate selection model allows genetic variation to be maintained in components of variation, trade-offs with monotonic curvature would lead to eventual loss of genetic variation in a wholly deterministic world because the curvature would lead to reduced fitness for any point away from the optimum on the curve. The multivariate selection model alone is therefore lacking as a full explanation of maintenance of variation. It is now accepted that recurrent mutation may play a role in maintaining quantitative variation in traits subject to normalizing selection. Following Lande (1975), a number of theoretical and simulation studies have lead to some agreement and some conflict (Rose et a l . , 1987; Lande, 1988; Turelli, 1988). Estimates of variance generated by new mutation in several traits in different X V E (Maynard Smith, species indicate results of the order of

I

I

Fig. 4 Constraints on life-history variables give rise to different equilibrium genetic correlations (Charlesworth, 1990): (a) and (e) show the pattern of genotypic variation for a trivariate system subject to constraints of the type used by Charlesworth (1990) to illustrate how both positive and genetic correlations can arise at genetic equilibrium; (b), (c) and (d) are the bivariate projections of (a) and all show negative genetic correlations; (f), (g) and (h) are the corresponding projections of (e) and, while two show negative genetic correlations, the genetic correlation between z2 and z3 at equilibrium is positive.

88

R. H . SMITH

1989a) and these values suggest that mutation may be more important than was once thought. However, Charlesworth (1990) found that mutations are unlikely to affect greatly the pattern of genetic correlations in a random-mating population. Mutation may contribute to observed variation in life-history characters but is otherwise of little immediate interest in ecological studies of life history and will not be considered further here. Non-additive effects are important (Rose, 1982, 1983b, 1985). A range of population genetics models with dominance, epistasis and other complications known to occur in real populations give rise to antagonistic pleiotropy, though there is little empirical evidence for the pattern of dominance required by the theory (Rose et a l . , 1987). Unfortunately the multivariate selection model does not easily generalize to include such interactions.

D. Gene-Environment Interactions As noted earlier, the basic quantitative genetics model assumes that there is no association between the genotypic effect g j k of an individual on a character z i , and the environmental effect ejk. If there is such interaction, the set of phenotypes produced by a genotype in different environments is known as the norm of reaction (Schmalhausen, 1949) of a genotype. The gene-environment interaction may mean that a particular difference in environment eik has a greater effect on some genotypes than on others, or there may be a change in the ranking of a series of genotypes measured on some phenotypic scale in different environments. For example, Falconer (1989a) compared the growth of young mice in two strains A and B under conditions of good and bad nutrition. In a good nutritional environment, strain A grew faster than strain B while in a bad nutritional environment, strain B grew faster than strain A. Note that this is a comparison of estimates of the mean value of ( r i + g i k ) in Eq. (lo), and that there could be an interaction of eik with either Z i (assumed constant within a strain) or gik (maybe variable within a strain) or both. Although the gene-environment interaction is logically regarded as part of the environmental variance (Falconer, 1989a, p. 135), because individuals experience different environments in nature, gene-environment interactions affect the realised phenotypes and hence the selection gradients for different genotypes. Indeed, Lewontin (1974) rejected simple partitioning of phenotypic variance into genetic and environmental components as a means of understanding why particular phenotypes have evolved, and argued that the norm of reaction is the real point of interest.

LIFE-HISTORY EVOLUTION

89

1. The Norm of Reaction and Adaptive Evolution Via and Lande (1985) and Via (1987) have discussed the norm of reaction as a potentially adaptive trait that can allow adaptive phenotypic plasticity (Bradshaw, 1965; Caswell, 1983; Moller et a l . , 1989b), and Stearns and Koella (1986) have developed an optimization model for a specific norm of reaction. Baker (1988) treated gene-environment interactions as individual responses to environmental stress and discussed their relevance to plant breeding. Defining and measuring a norm of reaction is not straightforward. The different environments (two or more) first have to be defined if gene-environment interactions are to be separated out in a structured account of phenotypic variance. Analysis of variance may then separate out gene-environment effects as a component of variance (Falconer, 1989a), but Via (1987) argued that this traditional approach is not very informative. An alternative approach is to treat a single character z , expressed in two environments e , and e2 as two character states (Falconer, 1952), each expressed in only one environment. The two states z , ( E , ) and z , ( E z ) may have some common genetic determination and hence there is a sort of genetic correlation ( I C E ) . Via (1987; Fig. 3) showed how i G E measures the gene-environment interaction. If the variance among genotypes V, is the same in each environment, the gene-environment interaction variance VGEis related to i G E as follows (Yamada, 1962):

In a designed breeding experiment, for example half-sib/full-sib families (see Section IV.C), ?-GE is estimated by the correlation of family means or breeding values (Falconer, 1989a) in the two character states. Interestingly, although r G E = - 1 represents the greatest gene-environment interactions, both positive and negative r G E constrain independent evolution of the two character states (Fig. 5 ) ; the greatest opportunity for phenotypic plasticity is when rGE = 0 (Via, 1987). For the multivariate case, Via and Lande (1985) noted that the evolution of a set of character states expressed in different environments depends critically upon the pattern of genetic correlations between states, rcE. Via and Lande (1985) extended Lande’s (1979, 1982) multivariate selection model to two environments using a matrix formulation incorporating I C E . Just as the change in a single character under natural selection is affected by selection on correlated characters (Section III.C), so the dynamics of evolution in each character state depend both on a direct response to selection in that environment and on correlated responses to selection in other environments experienced by the population (Via, 1987).

90

R . H . SMITH

2. Relevance to Ecologists? It is clear that the mathematical theory outlined in this section goes well beyond the data available to test the predictions. Ecologists will think in terms of, not just two or three, but many micro-environments, potentially giving rise to a vast array of rcE values that can never be estimated in practice. Because the environment experienced by an animal population varies both in space and in time, is there any value in separating VcE from V E and developing a theory of gene-environment interactions when the rate of change of environment experienced by different genotypes is so rapid compared with the potential rate of evolution of norms of reaction? There are two particular sorts of environmental variation that are likely to be important in this context. One is externally imposed stress due to environmental toxins; an example of this sort of variability is the range of host plants that might be available to a polyphagous insect, and this is the context in which Via and Lande (1985) developed the above theory (Via, 1984a, b). When larval stages cannot move from a host plant chosen for them by their mother from a range of options, phenotypic plasticity is clearly at a premium as a means of coping with plant secondary toxins even though there may be a cost of plasticity (and hence a trade-off) compared with obligate monophagy. Holloway et af ., (1990a,b) have examined character states in toxin-stressed environments by quantitative genetic analysis in the rice weevil Sitophifus oryzae, though they do not discuss their data in terms of gene-environment interactions but rather as effects of novel environments on genetic trade-offs (Service and Rose, 1985). The other environmental variation is internally imposed because of population pressure. Population density with its effects on I, and m, has been one of the main foci of attention of animal ecologists for decades, yet interactions of density with genetics have hardly been explored (Charlesworth, 1980). Many animals exploit patchily distributed resources, for example insect larvae that develop in seeds, fungi, carcases or dung. Even with adaptive oviposition behaviour to minimize adverse effects of larval competition (Smith and Lessells, 1985), the spatial distribution of patches of resources will result in patches at high and low larval density. Gene-density interactions may constrain the expression of genetic variation in, for example, body size and a correlated character at high density (see Section 1V.E). Holloway et al. (1990a) found that genetic correlations between fitness components were changed by larval density in S. oryzae, a species in which larval density (larvae/seed) is found t o vary substantially in a non-adaptive way (Smith and Lessells, 1985). The effect of high larval density on most species exploiting a patchy resource is to

91

(a)

Environment

E,

Environment

Environment

E:

E:

(b)

(c)

Fig. 5 The gene-environment (GE) interaction may take several forms. The phenotypic values of a character z 1 are plotted for several genotypes in two environments El and E z (after Via, 1987). By way of example, zl could be body weight at emergence in an insect at two population densities, high ( E 2 ) and low ( E l ) . In (a), there is no gene-environment interaction ( r G E = l), while in (b) the gene-environment interaction is so strong ( r G E = -1) that there is a complete reversal of ranked performance in the two environments. In (c), rGE = 0 because none of the genetic variation in z1 is expressed in the high-density environment Ez The gene-environment interaction in (c) means that selection would only act on z 1 in the low-density environment E 1.

reduce juvenile survivorship; thus the selection gradient will be changed (Eq. (17)), and there may be a change in the G matrix (effectively a gene-environment interaction). Estimation of genetic variances and covariances ought therefore to be carried out separately over a range of realistic densities, though in practice this is a mammoth task.

E. The G and E Matrices Section 1II.C concentrated on the genetic variances and covariances at equilibrium, given a selection differential VW and functional constraints

92

R. H . SMITH

fi. The rate of progress towards equilibrium also depends on non-genetic (environmental) variances and covariances. The multivariate analogue of Eq. (15) is (Lande, 1979): z = GP-‘S

where P - ’ is the inverse of the phenotypic variance-covariance matrix and S is the vector of selection differentials, typically defined in a truncation selection experiment as described in Section III.B.4. Thus the selection gradient VW is equivalent to the vector product P - ‘ S . Cheverud (1984) discussed the distinction between the selection gradient and the selection differential in relation to direct and indirect effects of selection. As discussed in Section IV.A, although phenotypic correlations can be quite different from genetic correlations (Cheverud, 1982), P and G will be very similar if heritabilities of the component traits are high (Bell and Koufopanou, 1986, p.104) because the multivariate analogue of Eq.( 11) is: P = G + E

(24)

As noted previously in relation to heritability, there are advantages in separating G from E and breaking these matrices down further where possible. Cheverud (1984) suggested that, because environmentally caused variations must act via the same development pathways as genetically caused variations, environmental correlations might be expected to show patterns similar to phenotypic correlations. Hegmann and DeFries (1970) suggested that environmental and genetic correlations are themselves correlated, though the available data show no clear pattern (Bell and Koufopanou, 1986, Fig. 9; Falconer, 1989a; Table 19.1). What use, if any, is the E matrix? Bell and Koufopanou (1986) noted that environmental correlation is regarded by geneticists as a nuisance because it obscures genetic correlation and reduces the selection response (Eq. (15)); they argued (pp. 105-108) that environmental correlations between components of fitness may be positive or negative if the components have different optima along an environmental gradient, or they may be positive or zero if the optima coincide. On this basis, the E matrix would seem to be just a nuisance! However, the E matrix must contain some information about physiological trade-offs within genotypes, that is developmental and resource-allocation constraints (Cheverud, 1984). The problem is how to separate the effects of resource allocation constraints from those of variation in resource acquisition (van Noordwijk and de Jong, 1986) and other effects such as gene-environment interaction. Individuals that are “lucky” for some non-genetic reason will tend to perform better in several components of

LIFE-HISTORY EVOLUTION

93

fitness than others that by chance are less fortunate. It is the confounding of effects of luck with those of constraints within individuals that make the simple E matrix hard to interpret (Table 4). Although there is as yet no clear solution to the problem of how to structure the E matrix into resource allocation vs resource acquisition components, some sort of manipulation experiment (Section 1V.B) within families in a designed breeding experiment (Section 1V.C) may be one answer. Experimental manipulations involving resource provision should mask chance variation in resource acquisition while designed breeding experiments take out genetic effects. Taking out gene-environment interactions as genetic covariation between character states (Via, 1987) should allow fundamental resource allocation constraints within individuals to be identified. Similar ideas to these are explored from a slightly different viewpoint by Stearns (1989).

IV. EXPERIMENTAL APPROACHES ILLUSTRATED BY CALLOSOBRUCHUS Life-history evolution can only be understood by identifying constraints-functional constraints that are related to development, physiology and resource allocation, and genetic constraints (Gould and Lewontin, 1979; Loeschke, 1987) that affect the potential of natural selection to drive the mean phenotype up the most direct route to an optimum (the selection gradient). Life-history trade-off constraints have been studied using a variety of methods (reviewed by Bell, 1980; Reznick, 1985; Bell and Koufopanou, 1986; Partridge and Harvey, 1988; Pease and Bull, 1988; Stearns, 1989). Reznick (1985) considered four methods (see below) but argued that only trade-offs with a genetic basis would constrain evolution, and therefore only the demonstration of additive genetic correlation o r of correlated response to selection were valid demonstrations of trade-offs. However, Bell and Koufopanou (1986) argued persuasively in favour of experimental manipulation, and Partridge and Harvey (1988) and Moller et al. (1989b) presented a case for carefully designed manipulations as means of measuring trade-off constraints. Only in a few studies have more than one of the four methods below been examined in the same population. The best-known demonstration of genetic trade-offs is the Drosophilu study by Rose and Charlesworth (1981a,b) who used a breeding experiment to estimate additive genetic correlations, and a selection experiment to evaluate the consequences; their results showed mainly negative correlations between components

94

R. H . SMITH

of fitness, though some were positive. Berven (1982) and Berven and Gill (1982) found additive genetic correlations indicative of a developmental constraint between developmental period and larval body size in the wood frog R a m sylvatica; however, they found a correlation between habitat and r A , which was positive in lowland and strongly negative in mountain populations. The milkweed bugs Oncopeltus spp. have been studied by Dingle and co-workers using both phenotypic and genetic methods (Dingle et af., 1982; Dingle, 1986; Scott and Dingle, 1990). While some authors argue strongly in favour of a particular approach to the exclusion of others, most agree that different approaches yield different sorts of complementary information (e.g. Stearns, 1989). Clearly, it is easier to work on a species that can be successfully reared in the laboratory, and more useful if the laboratory environment is near-identical or at least similar to some of the environments experienced by ancestral populations in the wild. That is why Drosophifu and pests of stored seeds are widely used, though we should always remember that concentrating on such species may give us an oversimplified and probably biased view of the natural world. In the brief outline of experimental approaches below, studies on the seed beetle C. rnaculatus will provide the main examples used for illustration.

A. Phenotypic Correlation Correlations between phenotypes may be carried out at a number of different levels: (i) (ii) (iii) (iv)

between species comparisons; between populations of the same species; between unspecified genotypes within species; between genetically identical members of a clone in species that reproduce asexually.

Although spurned by many, broad correlations across species generate most of the interesting questions about evolution of life-histories. Without such observations, strict experimentalists would not know where to start. There are problems with statistical analysis and interpretation (Clutton-Brock and Harvey, 1979). Perhaps the biggest problem is the widespread tendency to treat everything as a consequence of body size rather than question whether evolution of body size might not have been driven by other aspects of life history (Partridge and Harvey, 1988; Caswell, 1989a; Sibly et a [ . , 1991).

LIFE-HISTORY EVOLUTION

95

An example of the use of between-species comparisons to try to understand life-history evolution is the development of the idea of rand K-strategies. The conce?t of r- vs K-selection spawned a growth industry of studies of correlations across species in the 1970s, following Pianka’s (1970) expansion of the original idea. Weeds and pest species were generally described as r-strategists, and pest ecology is a field where between-species comparisons arise naturally because of the need to predict, understand and control the distribution of damaging pests. For example, Giga and Smith (1983) studied comparative aspects of life histories in four species of seed beetle, Callosobruchus spp. Some of their results will be described below because C. maculafus is used as an example in Sections 1V.B-E. Callosobruchus species are beetles of the family Bruchidae and are pests of stored legume seeds in the tropics and sub-tropics. Eggs are laid onto the surface of seeds or seed-pods, and the larvae develop within a single seed and can cause substantial loss of quality and quantity of seeds. Adult beetles do not normally feed in stores but may take nectar or similar sources of energy in the field. Giga and Smith (1983) estimated the following components of fitness in four Callosobruchus species under different environmental conditions in the laboratory: (i) development rate (ii) survivorship from egg to adult (iii) lifetime fecundity (iv) adult female longevity. Using an approximate expression (Howe, 1953), they estimated r for the different environmental conditions, and found that temperature (not unexpectedly) was the main determinant of r . The interest here in this standard approach is that it revealed a species-environment interaction, showing that of two closely related species, Callosobruchus rhodesianus was better adapted to cool conditions than C. macularus. The study also compared two geographical strains of C . maculatus, but these were hardly different (Table 1) although later studies (Credland et al., 1986) showed much greater divergence in other geographical strains of the same species. The life-cycle of the four Callosobruchus species is summarized in Table 1. Points to note from Giga and Smith (1983) are that all species are semelparous, though Callosobruchus analis has a greatly extended oviposition period and longevity compared with the other species, and a correspondingly reduced lifetime fecundity (see Section 1V.B for a physiological explanation); this tendency towards iteroparity is correlated with low juvenile survivorship, itself a consequence of intense larval competition (Smith and Lessells, 1985; Smith, 1990). Thus phenotypic correlations reveal a suite or syndrome of life-history correlates (Bradshaw, 1990) and suggest where effort should be directed in more focussed investigation of genetic and physiological constraints.

96

K . ti. SMITH

Table 1 Life-history of four Callosobruchus species. Callosobruchus species

C. analis

C. chinensis C. tnaculatus C.rhodesianus (a) Life-history at 25 "C

Life-history character Egg-adult developmental period (days) Egg-adult survivorship ( % ) Lifetime fecundity Oviposition period (days) Female longevity (days)

40.5

31.4

28.0

35.9

22.6 48.5 12.9 15.3

67.3 70.7 6.2 7.5

82.9 67.4 6.2 6.3

66.3 49.0 5.9 8.5

(b) Intrinsic rate of increase (r) per female per calendar month Temp

("C)

20 25 30

35

-

-

1.06 1.37 0.44

2.76 4.03 2.91

0.48 3.20 3.60 3.60

0.60 2.15 2.73 0.60

Avoid High

Avoid High

Avoid High

5

4

5

Scramble

Scramble

Scramble

(c) Behaviour and dynamics Characteristic

Larval competition strategy

Attack Oviposition site selectivity Very high Time to lay 90% of eggs at 11 25 "C (days) Populationdynamicoutcome Contest

Results for C. macuhtus are for a Brazilian (Campinas) strain. Females attach eggs to the surface of dry legume seeds of which the cowpea Vigna unguicu/ara is the most suitable. Hatched larvae tunnel into the seed where development to pupation and adult emergence is completed. When there are several larvae within a single seed, larval competition causes increased mortality (Smith and Lessells. 1985). Emerged adults mate and lay eggs without delay. Adults do not normally feed. (a) Mean values of life-history characters recorded at low density on cowpeas at 25 "C (after Giga and Smith, 1983); (b) effect of temperature on the intrinsic rates of increase per calendar month, estimated using Howe's (1953) approximation; (c) summary of oviposition behaviour and population dynamics (Smith and Lessells, 1985). The data show that C. macularus is the best adapted to high temperature and C. rhodesianus to low temperature; C. analis has a substantially lower rate of increase because of extreme larval competition (Smith and Lessells, 1985) and consequently greater selectivity of oviposition site, greater longevity and reduced fecundity.

LIFE-HISTORY EVOLUTION

97

B. Experimental Manipulation Mdler et al. (1989a) argued that manipulation experiments are useful in at least two ways. First, they reveal parts of the options set that other approaches cannot reach (for example, genetic analysis is restricted to the range of genetic variation present in a study population at one time; Partridge and Harvey, 1988). Second, they measure contingent responses of individuals to a range of immediate circumstances; if this plastic response of the phenotype increases fitness and has a genetic basis, it will be favoured by selection (Caswell, 1983). Reznick (1985) argued that phenotypic plasticity does not reflect the response of organisms to natural selection and that only genetical analyses are of evolutionary interest because only genetically coded options take part in the evolutionary process. Bell and Koufopanou (1986) provided strong counter-arguments, pointing out that in general an experimental approach is more powerful than an observational one, and that properly designed experiments have provided a reasonably consistent, empirical basis for the theory of trade-offs, in contrast with estimation of genetic correlations. Partridge et al. (1986, 1987) successfully investigated fitness costs of reproduction in female D. rnelunogusfer, though Partridge (1989a,b) added the cautionary note that it is not yet clear to what extent responses to manipulation are similar to genetic effects. Manipulative studies in the field were reviewed by Partridge and Harvey (1988). Not surprisingly, many involve clutch or brood manipulation in birds (Nur, 1988) and data analysis has required fairly sophisticated model fitting (Pettifor et a1 ., 1988).

1 . Types of Manipulation Bell and Koufopanou (1986) categorized manipulations in two ways. First, similar groups may be given different treatments, e.g. brood size (Nur, 1984, 1988). opportunity to oviposit (Mdler et al., 1989b), mating opportunities (Partridge and Farquhar, 1981). Second, different groups may receive the same treatment, e.g. predation (Koufopanou and Bell, 1984), temperature manipulation (Giga and Smith, 1983). Bell and Koufopanou (1986, p.88) discussed problems of logic with the first category, though in general they regard the manipulative approach as having shed most light on the hypothesized cost of reproduction trade-off. There is a fair measure of agreement that manipulation experiments can successfully reveal physiological constraints provided experiments are carefully designed to control confounding variables such as phenotypic quality and environmental variation (Partridge 1989a,b; Partridge and Harvey, 1985, 1988).

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2. Changing the Options Set? It is implicit in the manipulative approach that the experiment should be exploring the options set (Fig. 6a) rather than creating new ones (Fig. 6b,c). The trade-off hypothesis is that there exists an options set bounded by the trade-off curve, and that manipulating a change in one character leads to a shift in the balance of resources allocated to different functions (Calow and Sibly, 1983) and a change in another character within the options set (Fig. 6a). However, an experimental manipulation might create a new environment with a changed options set. Figure 6b shows how creation of new options sets could give a spurious similarity to a trade-off within an options set. In contrast, Fig.

2,

(c)

Fig. 6. The effect of different experimental manipulations may be to create new options sets as in (b) and (c) rather than to move the mean values of a pair of characters along the trade-off curve as in (a). The arrows point to mean values observed in an unmanipulated ancestral environment (A) and in two manipulations (B and C). In (b) the new options sets created give a spurious similarity to the trade-off hypothesis (a negative correlation between components of fitness), while in (c) the new options sets apparently contradict the trade-off hypothesis. Only in (a) are the manipulations valid (after Moiler et al., 1989b).

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6c shows how new environments could give a picture that apparently contradicts the trade-off hypothesis; this can often arise if a manipulation affects resource availability (cf. van Noordwijk and de Jong, 1986) and variation in resource availability can mask variation in resource allocation. Because new options sets have been created, neither situation (Fig. 6b or 6c) provides information about trade-offs within an options set (Calow and Sibly, 1983; Fig. 4). In practice, it may not be straightforward to decide which manipulations are appropriate.

3. Callosobruchus maculatus The cowpea weevil C. maculatus has a fairly simple, semelparous life-history, and the ancestral environment is well known and easily replicated in laboratory conditions. Accordingly, Maller et al. (1989b) were able to design manipulations of juveniles and adults separately that fell into three categories: Normal environmental variation; juvenile density and opportunities for oviposition by adult females were manipulated within the range likely to be encountered in culture conditions. (ii) Temperature variation; temperature was varied above (30 "C) and below (23 "C) the normal culture temperature (25 "C) experienced over the previous 32 generations in the laboratory. (iii) Nutritional variation; juvenile nutrition was varied by changing the host plant seed that larvae were reared in, and adult nutrition by providing unusual feeding and drinking opportunities (food and water are not normally available for adults in culture). (i)

The full results are given in Mdler et al. (1989b; Fig. 2 , Tables 3 and 4). The broad pattern of correlations in the different categories of manipulation is summarized in Table 2. The results are in accord with a priori predictions about options sets (Fig. 6), and caution against uncritical use of manipulations. The main problem in manipulation experiments is deciding which characters to measure (and how many), and has not been addressed by theory as it has for genetic correlations (Charlesworth, 1990). It is worth noting that the negative correlation found between lifetime fecundity and adult longevity in C. maculatus using manipulation (Table 2) demonstrated a cost of reproduction trade-off that was not revealed by genetic analysis (Mdler et a l . , 1989a) because of lack of genetic variation. The physiological mechanism underlying this trade-off is easily understood; adults do not normally have feeding opportunities in

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Table 2 Results of manipulative experiments in Caflosobruchus maculutus Characters examined

Manipulations (No.) Correlation

(5) Emergence weight (z4) a. Density vs development rate ( z 2 )b.Temperature (3) (2) (juvenile) c. Nutrition

-

(4) a. Density b.Temperature (3) (5) c. Nutrition

-

Lifetime fecundity ( z , ) vs longevity ( z 3 ) (adults)

-

+ -

+

Interpretation Physiological trade-off Spurious ‘trade-off‘ New options set Physiological trade off Spurious ‘trade-off‘ New options set

The table (after Moller el al.. 1989b) summarizes the effects of three types of experimental manipulation on pairs of life-history characters in juveniles and adults. The trade-off hypothesis predicts a negative correlation between the characters examined as a consequence of a resource-allocation decision. It is argued by M ~ l l e rer al., (1989b) that only the density manipulations represented variation within the normal environment of the beetle; the temperature and nutrition manipulations d o not test the trade-off hypothesis because they create new environments and altered options sets.

culture, and have to use reserves accumulated during larval development for both maintenance and reproduction. When oviposition opportunities are not available, adults live longer at the cost of producing less eggs, and therefore can exploit opportunities for oviposition as and when they arise (e.g. by transfer to fresh seeds; Sibly et a l . , 1991). Thus there is adaptive value in this phenotypic plasticity, and the energy reserves that a beetle has at emergence represent a genuine constraint (Smith and Lessells, 1985).

C. Breeding Designs Genetic variances and covariances are estimated from phenotypic data, structured in such away as to allow estimation of correlation between relatives (Falconer, 1989a). The choice of experimental designs is greater in most plant species than in most animals because of the wider opportunities for cloning, selfing etc. in plants (Lawrence, 1984). A weakness in quantitative genetic studies, emphasized by Bell and Koufopanou (1986), is that, unless they involve selection or measurement of character states in different environments, they are essentially correlational, even though they involve “treatments” (familial relationships) in the statistical sense. In this sense, they are in the same category as phenotypic correlation studies. Indeed, given sufficiently high heritabilities, genetic correlations can be inferred quite accurately from phenotypic correlations (Bell and Koufopanou, 1986, p. 103) using standard formulae (Falconer, 1989a, p.315).

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The basic problem in estimating genetic parameters is that of separating out the genotypic effects g,k in Eq. (9) (the breeding value of the kth individual; Falconer, 1989a, p. 117). The genetic correlation is the correlation between breeding values, and genetic variance and heritability of course relate directly to the g,. There are very many examples of breeding experiments that estimate these genetic parameters in the animal and plant breeding literature (Falconer, 1989a), and too many even in the ecological and evolutionary literature to review here. Wright (1978), Law (1979), Dingle and Hegmann (1982) and Rose (1983a) give reviews, and Roff (1990) has recently summarized quantitative genetic studies of insect life-cycles. Overall, it seems that genetic correlation studies have often not identified the anticipated trade-offs (Bell and Koufopanou, 1986). In designing breeding experiments to estimate genetic parameters, a few basic rules need to be observed by those investigating life-history evolution: the study environment should be the same as (or as similar as possible to) the ancestral environment of the study species (Service and Rose, 1985; Clark, 1987; Holloway et a l . , 1990); the study population should be large and outbred, else there are likely to be artefactual correlations (see Rose (1984a,b) for a criticism of experiments by Giesel (1979), Giesel and Zettler (1980) and Giesel ef al. (1982a,b) who used inbred populations); also, linkage disequilibrium in small populations might give rise to genetic correlations that are interpreted as apparent pleiotropy (Rose et a l . , 1990); sample sizes must be large if estimates are to have anything like enough precision for inferences to be made; because variances and covariances are being compared, the sample size needs to be about the square of what would be regarded as a reasonable sample size when comparing means by normal distribution theory, that is n = 900 rather than n = 30 (Rose and Charlesworth, 1981a; Mdler er a l . , 1989a); the accumulated wisdom of decades of experience in animal and plant breeding should be drawn upon (Falconer, 1989a; Thompson, 1989).

1. Choice of Design Different experimental designs incorporate different causal components of variance in the observational components of phenotypic variance (Falconer, 1989a, Ch. 9). Relationships examined in different designs

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compare offspring with parents, half-sibs and full-sibs. Causal components of variance that should not be included in estimates of V A and r A are interaction effects (especially dominance deviations V,) and environmental covariance (especially maternal effects). The design of choice because it avoids bias is the half-sib design. An example is the design used by Mgller et af. (1989a) and is summarized in Table 3. A full-sib comparison should be avoided because, although the estimates are apparently precise, they are biased to an unknown degree by interaction variance and covariance, as well as maternal effects. Offspring-parent regression (either on single parent or on mid-parental value) does not suffer much from interaction effects, but can be complicated by maternal effects (if regression is on the mother) and by the offspring being measured in an environment that is not identical to that of their parents. All the above implies that choices are possible. Ecologists will often have to be opportunistic and make observations of similarity between relatives according to whatever comes up. While sample sizes are rarely sufficiently large in field studies, some long-term studies have generated estimates of genetic parameters (e.g. van Noordwijk et af., 1980). Modern computational methods that allow optimal analysis of data from different sorts of familial comparison offer better possibilities for the future study of field populations (Meyer, 1989), though there may often be a problem of non-random mating (e.g. large may tend to mate with large) that could lead to bias.

2. Statistical Analysis and Estimation Full-sibhalf-sib designs are conventionally analysed by nested analysis of variance. Unbalanced data sets can now be analysed using a statistical package such as SAS (SAS Institute Inc., 1985) with appropriate procedures (GLM and NESTED). Computing standard errors requires a theoretical approximation (e.g. Becker, 1984), and it is implicit in most analyses that the distribution of eik about gik is approximately normal. Standard errors are not always computed because the expressions are felt to be unreliable (e.g. Rose and Charlesworth, 1981a); it often turns out that the estimated standard errors are about as large as the estimates of h2 and r A , even with large sample sizes. Ecologists find more disconcerting the cases where point estimates of h 2 and r A are outside the feasible range (0-1 for h2 and -1 to +1 for r A ) ! This arises because of the way that components of variance are estimated by subtraction in nested analysis of variance, and one consequence is that sample estimates of VA and V E are likely to be negatively correlated even though by definition the population values are not. Unbalanced designs are prone to give biased estimates of variances, and it is

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sometimes difficult to achieve statistical balance in breeding experiments on animals where fecundity and viability are variable. An alternative estimation procedure to theoretical approximations is a non-parametric resampling method such as the jackknife or the bootstrap (Efron, 1982); resampling methods avoid the usual assumptions such as independence of parameters, normality and small variances, and are feasible nowadays because of the ready availability of fast computers. However, there are still problems about choosing appropriate transformations, especially if hypothesis testing is intended. Arvesen and Schmitz (1970) and Holloway ec a l . , (1990b) have used jacknife estimates in the analysis of quantitative genetic data. The introduction of restricted maximum likelihood estimation (REML) and best linear unbiased prediction methods (Meyer, 1989), again made possible by advances in computer technology, make it possible to analyse data involving more than one generation (Kennedy and Sorensen, 1988). Two further advantages of REML programs are that unbalanced data sets can be dealt with, and point estimates and confidence intervals can be constrained within the feasible range (W. G. Hill, pers. comm.).

3. Callosobruchus maculatus Mdler ef al. (1989a) used a full-sibhalf-sib breeding design to estimate heritabilities of and genetic correlations between life-history characters in C. maculatus. Two potential trade-offs were selected for investigation. The first was a constraint on fecundity and development rate. Fecundity is correlated with body weight at emergence in the Brazilian strain of C. maculatus studied (Smith and Lessells, 1985). However, large body size might be achieved only at the cost of slow development. Since body weight is likely to be the physiological link between fecundity and development rate, body weight at emergence was recorded as well as egg-to-adult development rate and fecundity. The second trade-off examined was between lifetime fecundity and adult longevity. As described in Section IV.B, adults have a finite amount of resource to allocate between fecundity and maintenance. Thus highly fecund genotypes may not be able to live as long as less fecund ones. In iteroparous species where the adults feed, the constraint would be more complicated. Moller et al. (1989a) analysed data on four measurements made on 761 offspring of 30 males, each mated to six females. Development rate had negative genetic correlations with both adult longevity and fecundity (Table 4), while longevity and fecundity had a positive genetic correlation. Being larger genetically increases the fitness of adults in two ways, by increasing both fecundity and longevity. In order to remove the joint

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dependence on weight, residual fecundity and residual adult longevity were calculated for each individual after removing the effects of weight by linear regression. The genetic correlation between residual fecundity and residual adult longevity was not significantly different from zero, indicating that there was no genetic trade-off, though experimental manipulation did reveal a physiological trade-off (Mdler et al., 1989b). Overall, the results did show some genetic trade-offs, and emphasized the central role of body weight as an apparently causal link between components of fitness. Messina (1989, 1990) has also examined quantitative genetic parameters in C . macufarus, though his results cannot be compared directly because he measured different variables in a different population. The paper by Charlesworth (1990) now provides expressions that allow further analysis of multivariate genetic data. For C. macufarus, the estimated values of V, and VE can be calculated from the published values of r p , rA and h2 (Mgiller et a f . , 1989a; Table 4) and of V,, (Mplller et a f . , 1989a; Table 3, where the standard deviations are mis-labelled as standard errors; H. Moller, pers. comm.). The estimates of the causal components of variance are given in Table 3 along with the coefficient of genetic variation, which ranged from 3% for development rate to 12% for fecundity (cf. comparable figures for D . mefanogaster in Charlesworth, 1984). The phenotypic, genetic and environmental variance-covariance matrices are given in Table 4, with the corresponding correlations in the lower half of each tabulated matrix. The figures are given here to two decimal places without any measure of precision. Using the estimated G matrix (Table 4b), it is possible to test whether the estimated figures indicate genetic equilibrium or not. For example, Eq. (14) of Charlesworth (1990) for the covariance between constrained variables as a function of other variances and covariances is:

For the covariance between fecundity (zl) and development rate ( z 2 ) which were suggested a priori to be constrained, Ct2 should be approximately (C13x C23)/V33+ (C14 x C24)/V44, if weight and adult longevity are treated as free variables. The predicted C12 is -4.18, which is within 7% of the observed estimate (-3.90); this good agreement supports the premise that the estimated G matrix represents equilibrium variances and covariances, and perhaps suggests that the totality of the estimated genetic correlation matrix is more reliable than the approximate standard errors of individual point estimates suggest. Note that if weight had not been measured and included in the analysis, the predicted value of C12would have been -2.04, barely 52%

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LIFE-HISTORY EVOLUTION

of the observed estimate. This underlines the importance of measuring all the variables involved in the genetic constraints (Pease and Bull, 1988; Charlesworth, 1990) in order to be able to understand the pattern of genetic correlations and emphasizes the inadequacy of a simple, bivariate approach to genetic trade-offs. Calculation of the determinant of G and of the determinants of sub-sets of G with different variables removed makes the same point (the determinant of G should be close to zero at equilibrium). Table 3 Quantitative genetic analysis of the life-history of Cullosobruchus muculutus

(a) Breeding design Male parent

Female parent

Female offspring

Offspring characters recorded

1

2

E.1

(b) Phenotypic means, heritabilities and variances Life-history character Mean

z1 Fecundity (eggs) z 2 Development rate ( x lo3 days-') 23 Adult longevity (days) 24 Weight at emergence (g)

Heritability

Z

h2

82 34

Variances

v,

v,

0.10 0.23

104 4

10

94

0.124

0.9

3.1

0.028

8.2

0.30

1.0

0.3

0.7

0.067

5.2

0.55

0.22

0.12

0.10

0.067

v,

vyyz

(After Moiler er a / . , 1989a.) (a) The full-sibhalf-sib design used to estimate genetic and environmental components of variances and covariances. In the hierarchical design, 30 males were mated with six females each, and four life-history characters were recorded for each of five female offspring per female parent. (b) Summary of narrow-sense heritability estimates, taken from Mdler er al. (1989a). The additive genetic (V,) and environmental (V,) components of phenotypic (V,) variance were calculated using the standard deviations of the mean phenotypic values in Table 3 of Mdler et al. (1989a). The quantity V y z / i varies from 3 to 12% and is a relative measure of the genetic contribution to the mean phenotype.

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R. H. SMITH

Table 4 The phenotypic, genetic and environmental variance-covariance matrices estimated for the reduced data set (761 offspring) of Moller et al., (1989a). In each case, the variances (italicized) are on the diagonal, the covariances are above the diagonal, and the correlations are below the diagonal

z1

z2

z3

7-4

104.04 -0.04 +0.43 +0.61

-0.82 4.00 +0.06 -0.02

+4.39 +0.12 1.00 +0.33

+2.92 -0.02 +O. 16 0.22

10.40 +0.91 +0.68

-3.90 0.92 -0.73 - 1.01

+1.61 -0.38 0.30 +0.76

-0.76 -0.34 +O. 15 0.12

93.64 +0.18 +0.34 t-0.71

+3.08 3.08 +0.34 -0.58

+2.78 +0.50 0.70 +0.04

+2.16 +0.32

(a) P Matrix ZI 22 z3 z4

(b) G Matrix 21

22 z3 z4

- 1.26

(c) E Matrix

+0.01 0.10

One of the most important results of multivariate selection theory highlighted by Pease and Bull (1988) and Charlesworth (1990) is illustrated by the G matrix of C. mucufurus and should be emphasized once again: at equilibrium, a positive genetic correlation may evolve between two components of fitness even though the underlying constraints in the whole set of components of fitness are negative constraints of the trade-off type. In C. macufutus, fecundity and adult longevity were positively genetically correlated (Table 4b) although the underlying physiological trade-off constraint reflecting resource allocation was negative (Table 2).

D. Selection Experiments Changing the selection gradient should lead to observable changes in mean values of ti(and possibly in the G matrix). If genetic correlations are real, direct selection on a variable may be constrained and there should be a correlated response to selection in other variables. Bell and Koufopanou (1986) argued in favour of selection experiments rather

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than breeding designs because a correlated response to selection demonstrates additive genetic correlation without ambiguity, and selection experiments seem to have demonstrated trade-offs more consistently. Rose and Charlesworth (1981a) found a correlated response to selection for early and late life-history characters in D . melanogaster, and Luckinbill et al. (1984) and Rose (1984b) confirmed these results.

1. Callosobruchus maculatus Having demonstrated additive genetic variation and covariation in C. maculatus, Mqjller et al. (1990) reported on a laboratory natural selection experiment where selection was exercised by using a culture interval (the time between transferring adult insects onto a fresh set of seeds) that was either very short or very long. Individuals that emerge just before transfer will have highest fitness. Those that emerge too late for transfer perish, while those that emerge too early miss the opportunity for additional growth (and fecundity), use up resources in staying alive, and perhaps waste eggs on old seeds that will be discarded. Thus short culture intervals favour fast developers which suffer a cost of light body weight, low lifetime fecundity and reduced adult longevity. In contrast, long culture intervals favour slow developers with heavy body weight, high lifetime fecundity and increased adult longevity. Each selection regime was set up as four replicate lines, and the short culture interval (27 day) regime underwent five generations of selection while the long culture interval (38 day) regime underwent four generations. To eliminate non-genetical (maternal) effects, selection was relaxed for two generations in all lines before testing (Mgller et al., 1990). The results of the selection experiment were qualitatively as predicted. Curiously, development rate (the character under direct selection) was not measured, but weight at emergence and adult longevity were. Weight at emergence was significantly higher in the long culture interval regime, while adult longevity was correspondingly higher also (Mdler et al., 1990). Did the correlated response to selection fit with quantitative predictions based on the genetic variances and covariances? From Eq. (16), direct selection on a character zi (development rate) should lead to a correlated response in another character z j , the magnitude of which depends on the genetic correlation and the additive genetic variances. For a third character z k , the correlated response can also be predicted. With direct selection on z i , the ratio of correlated responses in z j and z k is : Rj/Rk

If

ti

=

(rA(ir

is longevity ( z 3 ) and

j)/rA(i,

zk

k))(VA(zj)/VA(Zk))1’2

(26)

is weight ( z 4 ) , the expected indirect

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R. H . SMITH

responses to direct selection on development rate are related by:

(ZJ

in C. macularus

R3/R4 = (-1.01/-0.73)(0.30/0.12)”2 2.17 (27) The observed responses to selection can be summarized as the difference between the means in the two selection environments divided by their mean: for longevity z 3 , this was calculated from the selection regime mean longevities (10.40 - 10.87)/10.635 (from Table 2a of Moiler et a l . , 1990), while for weight 24 it was (2.369 - 2.652)/2.51. The observed selection responses then were R3 = -4.4% and R4 = -11.3%, giving R3/R4 = 2.55; thus the predicted selection response ratio was 15% less than the observed, which is not too bad a margin of error given the number of parameter estimates involved in the calculation (again suggesting that the point estimates might be more precise than indicated by their approximate standard errors). It appears, then, that for the two characters measured, the results of short-term indirect selection correspond well with the predictions of a much more laborious breeding experiment. However, it is not possible to test for genetic equilibrium using the results of a selection experiment which do not estimate the G matrix, and the G matrix will change under selection. Ideally, the G matrix would be estimated at intervals during the course of selection, though in practice this does not seem feasible. This quantitative check on predictions is an alternative to the graphical approach used by Smith et al. (1987) for the results of selection on P. rapae that looked at how selection affected fitness. Using a theoretical expression for fitness in P. rapae, Smith et al. (1987) found that selecting upwards on a component of fitness only had a small effect on mean fitness because of a correlated response in another component. Selecting downwards lead to a reduction in both components of fitness, suggesting that there were maladaptive genotypes present that might be maintained by recurrent mutation (Charlesworth, 1990). =

2. Practical Difficulties Selection experiments, however, are not always straightforward, and for practical reasons they tend to be restricted to laboratory population studies. Reznick (1985) noted some problems; for example, it is more difficult to control the environment for the duration of a selection experiment than it is for a two-generation breeding design (Steams, 1989). Rose et al. (1990) reviewed the use of selection to probe patterns of pleiotropy, and described a number of alternative types of experimental design. For reasons of practicality and maintaining sufficiently

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109

large populations to overcome linkage disequilibrium, laboratory natural selection experiments are preferred to truncation selection or other procedures that are more interventionist (Rose et al., 1990). Mueller (1987) and Rose et al. (1990) documented results of selection experiments in D . rnelanoguster and concluded that negative pleiotropy is important, but not universally important, in moulding life history in Drosophila.

E. Mapping the Options Set To what extent can it be said that the above approaches map the options set, or even delineate the boundary of the options set (the trade-off curve)? There is clearly a distinction between the genetic options set in a particular environment (mapped out by plotting estimated breeding values-Smith et al., 1987; Mdler et al., 1989a) and the phenotypic options set which encompasses physiological trade-offs and non-adaptive random effects on top of the genotypic values. A combination of manipulations and breeding designs would in theory estimate both genetic and physiological trade-offs (Mplller et al., 1989b) another way of describing gene-environment interactions. In practice, selection in different manipulated environments is likely to be the most efficient way to proceed. Genetic and physiological trade-offs may produce some surprizing constraints on characters that are not directly related to fitness. For example, Wrelton (1987) and Li (1988) following Roff (1981) have produced phenotypic models of optimal body size in the beetles S. orytue and Prostephanus truncatus based on observed phenotypic correlations between body size and components of fitness. Since at least some of the characters concerned had high heritability, their models probably reflect underlying genetic trade-offs that indirectly stabilize body weight. Genetic and phenotypic correlations have been for the first time combined by Sibly et al. (1991) in a model of fitness using published data on C. rnaculatus (Mdler et al., 1989a,b). In Sibly er al. (1991), an explicit model of fitness in culture was developed based on the primary components of fitness (fecundity, development rate, longevity, oviposition behaviour). The genetic regressions of fecundity and development rate on weight were then put into the model, and it was found that these genetic constraints produced indirect stabilizing selection on female body weight, with a predicted optimal body weight corresponding to the range found in culture. Culture interval was found to be a key environmental variable (Fig. 7). This integrative approach is one that might be of value in other studies, though it depends on being able to write down a realistic expression for fitness in a particular species (e.g.

110

R. H . SMITH

Smith et a l . , 1987; Sibly et a l . , 1991). Certainly, the studies on C. rnaculatus have revealed the complementary nature of the four different experimental approaches outlined above. Although the C matrix at equilibrium reflects underlying functional constraints (Charlesworth, 1990), the complex nature of the multivariate case means that G is a veritable hall of mirrors, and examining only one part of it can give a distorted view of the whole.

V. DISCUSSION Rose et al., (1987) reviewed three approaches to trade-offs in life-history evolution and questioned whether optimization theory was wellfounded at the population genetic level; they suggested that optimization theory is simply outmoded in the study of life-history evolution. It offers little in the way of predictions about evolutionary mechanisms, and still less does it suggest fruitful avenues for further experimentation. From this point on, we expect it to become a mere distraction, at least where research on life-history evolution is concerned. Other topics within evolutionary biology might be illuminated by the pale light that it casts, but life-history evolution is no longer one of those.

In the context of this review, it is important to consider whether or not the strong views of Rose et al. (1987) are justified. The answer must surely be no. The recent results of Charnov (1989) and Charlesworth

0.10-

-

r

6

2 005L

O 0

I

I

4

8

Body weight ( rng )

Fig. 7. Indirect stabilizing selection on female body weight at emergence in C. macularus. Fecundity (z I ) and development rate (z2) are respectively positively and negatively genetically correlated with body weight (24). An expression for fitness ( r ) in laboratory culture based on genetic regressions of z1 and z2 on z4 is plotted against body weight. Culture interval (25, 30 or 35 days) strongly affects optimal body size (after Sibly el al., 1991).

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111

(1990) show a clear link between the results of optimization theory and those of evolutionary quantitative genetics. It could even be argued that strict population genetics has only a limited amount to offer life-history evolution. While particular population genetic models (e .g. Rose, 1982) have shown how, for example, a pattern of pleiotropy and dominance can maintain polymorphism, it is difficult to generalize from these special cases. Yet we find that the optimality approach, for example, with an explicit assumption of a single dominant gene determining life-history (Sibly and Calow, 1986, p . l l ) , gives the same prediction about mean life-history at equilibrium as the multivariate quantitative genetic approach with the explicit assumptions of additive effects (no dominance) and weak selection (Charlesworth, 1990); this gives us some confidence in the validity of the results. Furthermore, as shown here by development of a general iteroparous life-history model (Eq. (3)), the properties of the G matrix at equilibrium bear a very close relationship to the optimality criterion. Generalization to the case of frequency-dependent selection reveals again that frequency-dependent optimality results (an ESS solution) are the same as the predicted mean from multivariate quantitative genetics (Charlesworth, 1990). As noted by Maynard Smith (1989b), classical population genetics has had little to say about frequency-dependent selection. Once again, there is a correspondence between the results of an explicit genetic model with the assumption of dominance and the ESS predictions of the game-theory approach to frequency-dependent optimization (e.g. Smith, 1990). The classical problem of frequencydependent selection is evolution of the sex ratio (Fisher, 1930)-a problem in life-history evolution that has been illuminated most clearly by optimization theory (Charnov, 1982). However, the results of Pease and Bull (1988) and Charlesworth (1990) emphasize that multivariate selection may be characterized by a pattern of genetic constraints that cannot be predicted from considering functional constraints between variables taken two at a time, as is the usual approach in optimization theory (Sibly and Calow, 1983) as well as in discussions of negative pleiotropy (Rose, 1983a; Smith et al., 1987). If some components of fitness are physiologically linked to characters such as body weight (e.g. the C. macularus example, Mdler er al., 1989a,b; Sibly et al., 1991), then the pattern of genetic variances and covariances cannot be understood without inclusion of what are, in effect, constraining physiological and developmental variables (Cheverud, 1984). This brings us back to the value of experimental manipulations in helping to understand life-history evolution (Bell and Koufopanou, 1986). As noted by Maynard Smith (1989b) and Parker and Maynard Smith (1990), the G matrix may be important and useful (though not a

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constant in evolution), but on its own it has limited utility; the best known genetic trade-off, that between longevity and fecundity in Drosophila (Rose and Charlesworth, 1981a) arose from physiological constraints whose existence was well established experimentally, and which would probably be hard to alter by selection, although they are not understood at a biochemical level. (Maynard Smith, 1989b, p. 72).

Parker and Maynard Smith (1990) argued that physiological experiments and interspecies comparisons may be easier than quantitative genetics as means of discovering such constraints. Finally, it is important to emphasize the potential role of environmental patchiness in life-history evolution (Shorrocks, 1990). One effect of variation in patch quality with gene flow between patches is that it may not be possible to evolve towards a life-history that is optimal for each patch (Dhondt et a [ . , 1990). There may also be consequences of frequency-dependent selection (Parker and Maynard Smith, 1990). The effects on population dynamics of uneven distribution of animals between, and hence uneven competition within, patches are now well known (Shorrocks, 1990). In the context of life-history evolution, spatial heterogeneity is likely to lead to frequency-dependent selection on alternative strategies of competition between juveniles, and adapted suites of correlated life-history characters that evolve along with different strategies (Table 1): a prime example of this is seed beetles with internally feeding larvae, whose competition strategies and population dynamics can be broadly described as “scramble” or “contest” but reflect frequency dependence at the level of larval behaviour (Smith and Lessells, 1985; Smith 1990). However, there are many other species with mobile adults whose larvae are confined to patches of resource (e.g. insect parasitoids, dung flies, carrion flies and many drosophilids feeding on fruit and fungi) that may also be subject to frequency-dependent selection affecting juvenile survivorship and hence other life-history characters. One consequence of this may be the existence in nature of populations with alternative ESS life-histories characterized by different equilibrium G matrices, a concept that bears some relationship to the notion of adaptive peaks in Wright’s (1931) shifting balance theory of evolution and could be relevant to speciation.

ACKNOWLEDGEMENTS The synthesis attempted here has been greatly helped by discussions with a number of colleagues, and especially Richard Sibly. In particular, our understanding of some of the multivariate G matrix results of

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Charlesworth (1990) was enhanced by Richard Sibly’s modelling skills in this case with cardboard and sticky tape, though his use of scissors added a new dimension t o the phrase “linear approximation”. I have made extensive reference t o the published results of practical work that Henrik Maller did in my laboratory during his year in Reading in 1984-85 supported by the British Council. For comments on the text I thank Robert Curnow, Guy Ovenden and Richard Sibley. Finally, I am very grateful t o Michael Begon for his patience and encouragement, and his persuasive coaxing t o try t o keep me t o a deadline.

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Ecological Implications of Specificity Between Plants and Rhizosphere Micro-organisms C . P . CHANWAY. R . TURKINGTON. and F . B . HOLL

.................................... I11. Rhizohilrr?? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A . Infectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B . Effectiveness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1. Introduction

I1 . Specificity Between Plants and Beneficial Micro-organisms . . . . . .

C . The Relationship of Infectivity to Effectiveness . . . . . . . . IV . Frntikia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A . Infectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B . Effectiveness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V . Associative Rhizosphere Bacteria . . . . . . . . . . . . . . . . . . . . . . A . Infectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B . Effectiveness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VI . Ectomycorrhizal Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A . Infectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B . Effectiveness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VII . Vesicular-Arbuscular Mycorrhizal (VAM) Fungi . . . . . . . . . . . . A . Infectivity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B . Effectiveness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . VIII . Effects on Plant Competition . . . . . . . . . . . . . . . . . . . . . . . . . A . Symbiotic Bacteria . . . . . . . . . . . . . . . . . . . . . . . . . . . B . Associative Rhizosphere Bacteria . . . . . . . . . . . . . . . . . C . Mycorrhizal Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . IX . Consequences of Specificity on Plant Community Structure . . . . . X . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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I. INTRODUCTION The structure of plant communities is influenced by the interactions among its component species and associated fauna. Conventionally, the most important interactions are competition and herbivory (Harper, 1977; Grace and Tilman, 1990). The intensity and direction of these interactions is influenced by abiotic factors and by other, biotic, factors which have not been well studied e.g. micro-organisms, pathogens, and mycorrhizal fungi. It is not yet fully understood what direct impact these factors have on the outcome of competition and consequently on community structure, but preliminary evidence suggests that it could be substantial. The outcome of “competition experiments” can be altered, or even reversed by mycorrhizas (Fitter, 1977), Rhizobiurn strains (Young and Mytton, 1983; Mytton and Hughes, 1984; Chanway et a l . , 1989a), parasites (Dobson and Hudson, 1986) and viruses (Mackenzie, 1985). These organisms may negate competition by preventing dominance, exclusion, or the establishment of equilibrium conditions, and will, therefore, have an impact on winners and losers. Ultimately, they will influence which genotypes of which species continue to play the evolutionary game. The influence of micro-organisms on interactions between plants, and on the outcome of competition has received little attention, at least in part, due to the technical difficulties that plant population biologists may encounter with microbiological techniques. However, there can be considerable diversity in the population of rhizosphere micro-organisms and in their effects on plant growth. Hiltner (1904) was the first to recognize the importance of microbial activity associated with root systems in plant nutrition and used the term “rhizosphere” to describe this zone of intense microbial activity around roots of the Leguminosae (Fabaceae) (see Table 1 for definition of terms used in this review). Several genera of bacteria, fungi, viruses, and soil fauna are commonly represented (Richards, 1987). The influence of rhizosphere micro-organisms on plant growth and plant competitive ability in natural plant communities is substantial and biotic interactions between roots and associated micro-organisms can result in positive or negative impacts on plant productivity (Newman, 1978; Gaskins ef al., 1985; Curl and Truelove, 1986; Turkington et al., 1988; Chanway et a l . , 1988a,b, 1989a,b, 1990). Measurement of microbial biomass or population size in soil is difficult, and the techniques commonly used have limited accuracy. For example, direct microscopic examination of soil samples tends to overestimate the number of viable micro-organisms because non-viable cells

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Table 1 Definition of terms used in this review Associative :

Description of an association between dissimilar living organisms in which intimate physical contact or attachment is lacking but mutual benefits are realized e.g. PGPR.

Bacteroid:

Legume root nodule bacterium that has morphologically and physiologically differentiated into a cell that is capable of symbiotic nitrogen fixation within the nodule.

Conjugation:

Transfer of D N A from a bacterial donor cell to a recipient cell involving direct cellular contact.

Ectom ycorrhiza:

A symbiotic association between a fungus and plant root characterized by formation of a relatively thick mantle by fungal hyphae and penetration of mycelial strands inward between cortical cells and outward into the soil.

Endomycorrhiza :

A symbiotic association between a fungus and plant root characterized by extensive inter- and intracellular penetration by fungal hyphae, but with no mantle formation, and by mycelial penetration outward into the soil.

Mucigel:

Gelatinous material secreted by plant roots which is comprised of simple and complex carbohydrates.

PGPR:

Plant Growth-Promoting Rhizobacteria comprise a group of naturally occurring soil bacteria that colonize root systems and stimulate plant growth.

Rhizoplane:

The root surface.

Rhizosheath:

Soil which tightly adheres to the rhizoplane.

Rhizosphere :

The region of soil surrounding plant roots and influenced by their metabolism.

Sporocarp:

A hard, usually globose multicellular fruiting body that contains spores e.g. mushrooms.

Symbiotic:

Description of a close association between two dissimilar living organisms often involving physical attachment and generally assumed to be mutually beneficial e.g. nitrogenfixing root-nodule bacteria.

Transduction:

Transfer of a D N A fragment from a bacterial donor cell to a recipient cell mediated by a bacterial virus.

Transformation:

Uptake of a D N A fragment by a bacterial recipient cell from growth medium which had been previously released by a donor cell.

are included in counts (Russell, 1973). Alternatively, indirect “plate count” techniques have been used to estimate soil micro-organism populations. A known weight of soil is suspended in a buffer, diluted several times, and an aliquot of each dilution is poured into Petri dishes

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containing a nutrient agar. After incubation, the number of colonies growing on the plate is counted, and related to the weight of the original soil sample. Due to factors such as incompatibility between microbial propagules in the soil sample and the nutrient medium used for incubation, or incomplete suspension of microbial propagules in the original soil dilution, plate count techniques tend to underestimate the number of soil micro-organisms (Russell, 1973). A recently developed molecular technique based on the analysis of ribosomal ribonucleic acid (RNA) offers a culture-independent method; for studying microbial ecology and should result in a more objective assessment of variability and population size in natural microbial communities (Ward et af., 1990). Notwithstanding problems of measurement, Griffin (1972) estimated that field soil may contain 10'-10* bacterial cells c C 3 . Russell (1973) calculated the live bacterial biomass in the top 15 cm of Rothamsted field soils (Barnfield and Park Grass plots) to be between 1500 and 3500 kg ha-'. Jenkinson and Ladd (1981) tabulated direct measurements of microbial biomass in a range of soils and found that up to 3710 pg of carbon g-' soil was attributable to soil micro-organisms. The ratio of microbial numbers per unit mass of rhizosphere soil ( R ) to the numbers per unit mass of non-rhizosphere soil (S) is typically 10-5O:l for bacteria and 5-10:l for fungi (Richards, 1987). However, due to problems in deliniating R soil from S soil, and limitations associated with measuring the number of micro-organisms in soil samples, these ratios are approximate. Bacteria comprise the most common class of rhizosphere microbe (Rovira and Davey, 1974) and attain rhizosphere populations of up to 3 x 10" cellsg-' of soil (Rouatt and Katznelson, 1961). Because nutrients obtained by the plant root from the bulk soil must pass through the rhizosphere, microbial alteration of plant nutrients may be expected to affect plant growth. Several categories of rhizosphere micro-organisms are currently recognized (Table 2), including those which elicit stimulatory (Gaskins ef a l . , 1985) and inhibitory (Schroth and Weinhold, 1986) responses in the growth of the plant host. Some micro-organisms can reduce plant growth or inhibit reproductive yield without causing visible symptoms of disease. Bacteria that have this effect on plant growth have been termed deleterious rhizobacteria (Suslow and Schroth, 1982). Plant growth-promoting micro-organisms have been studied intensively because of their potential impact on agricultural and forest productivity. This group comprises symbiotic and associative bacteria and fungi and is the focus of this review. (The term "symbiosis" will be used as defined by Law and Lewis (1983) and will include only

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Table 2 Major categories of rhizosphere micro-organisms capable of altering plant growth Category Disease-causing pathogens Deleterious rhizobacteria Plant growth-promoting rhizobacteria Symbiotic bacteria (root nodule bacteria) Symbiotic (mycorrhizal) fungi

Reference Curl and Truelove (1986) Suslow and Schroth (1982) Kloepper ec al. (1980) Postgate (1982) Harley and Smith (1983)

mutualistic interactions between dissimilar organisms, not parasitism or commensalism as originally proposed by de Bary (1887)). The distinction between symbiotic and associative plant-microbe relationships is not always clear. Generally, when physical attachment of the micro-organism to the plant host occurs (e.g. formation of a root nodule by Rhizobium or infection of a root tip by a mycorrhizal fungus) the relationship is considered to be symbiotic. Bacteria that proliferate in the rhizosphere with no obvious sign of attachment to the root system or formation of special organs are associative. In both cases, mutual benefit to the host plant and associated micro-oganism(s) is assumed to occur. There has been renewed interest in the use of plant growth-promoting rhizobacteria (PGPR) (Suslow, 1982; Gaskins et a l . , 1985) since initial large-scale experimentation was performed by Russian agronomists in the 1950s (Cooper, 1959; Mishustin, 1963). The PGPR are naturally occurring free-living soil bacteria that are capable of colonizing roots and enhancing plant growth when added to seeds, roots, or tubers (Kloepper et a l . , 1980). There are several mechanisms by which PGPR may stimulate plant growth (Table 3) (see Lynch, 1988; Kloepper er al., 1989 for reviews), and the debate as to their relative importance is ongoing. Resolution of this problem is difficult because many species of PGPR possess attributes that are consistent with several of the proposed mechanisms (Holl et a l . , 1988). The most studied symbiotic bacteria belong to the genera Rhizobium and Bradyrhizobium. These root nodule bacteria have been studied for decades because of the substantial amounts of atmospheric nitrogen that may be fixed (up to 300 kgha-' year-') when in symbiosis with the appropriate legume host (Postgate, 1982). A small number of non-leguminous plants form nitrogen fixing root nodules after infection with filamentous bacteria belonging to the genus Frankia. This so-called actinorhizal symbiosis, exemplified by the nodulation of alder ( A l m s ) , is analogous to the Rhizobium-legume system but appears to be less evolutionarily advanced and less effective in nitrogen fixation. As our

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Table 3

Proposed mechanisms of action for plant growth-promoting rhizobacteria (PGPR) Mechanistn

Effect on plant growth

Reference

Phosphorus

Increased biomass and phosphorus content

Gerretsen (1948); Kundu and Gaur (1980); Subba Rao (1982) Kloepper and Schroth (1981); Suslow and Schroth

solubilization

Incrcased biomass Suppression of deleterious rhizobacteria Production of plant Increased shoot or root growth substances biomass; increased root branching: induce reproductive cycle Increased biomass and Root-associated nitrogen content nitrogen fixation

(1982)

Brown et al. (1968); Brown and Burlingham (1968); Brown (1974); Tien et al.

(1979) Dobereiner and Campelo (1971); Dobereiner et al. (1972); Rennie and Larson (1979); Rennie and Thomas (1987)

knowledge of the actinorhizal symbiosis increases, these generalizations may require adjustment. Mycorrhizal fungi are symbiotic root-infecting fungi that have been broadly categorized into one of two groups. Ectomycorrhizas are commonly associated with forest trees, including the families Pinaceae (pine), Betulaceae (birch), and Fagaceae (beech) in temperate regions. Colonization of plant roots results in formation of a morphologically distinct mycorrhiza (literal meaning “fungus root”), which is characterized by a dense fungal sheath surrounding the surface of the infected root (the mantle) and proliferation of the fungal mycelium between cells of the root cortex to form the Hartig net. Little or no intracellular penetration of cortical cells occurs, but root morphology is often altered. Endomycorrhizas are commonly associated with almost all taxonomic groups of plants including several angiosperm families of great economic importance such as the legumes and grasses. These fungi do not form an extensive sheath or a Hartig net, but intracellular penetration of cortical cells is common. Vesicular-arbuscular mycorrhizas (VAM) are the most abundant type of endomycorrhiza in nature, and are so named because of the specialized organs (vesicles and arbuscules) they form after penetration of cortical cells. Other categories of endomycorrhizas include arbutoid, monotropoid, ericoid, and orchid mycorrhizas. Harley

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and Smith (1983) have provided a comprehensive treatment of the biology of mycorrhizas.

11. SPECIFICITY BETWEEN PLANTS AND BENEFICIAL MICRO-ORGANISMS Root exuduates vary qualititively and quantitatively beween plant species (Rovira and Davey. 1974; Curl and Truelove, 1986) and between cultivars and genotypes of the same species (Keeling, 1974; Kraft, 1974; Baldani and Dobereiner, 1980). Because root associated micro-organisms (symbiotic or associative) depend on root exudates for nutrition and growth (Rovira, 1969), differences in root exudation can affect the composition of rhizosphere microbe populations and may result in microflora specific to plant species and plant genotypes (Neal et a l . , 1970, 1973). Specificity between plants and growth-promoting micro-organisms can occur at either of two stages of these associations; during infection of the root system to form root nodules or mycorrhizas, or during subsequent growth of the infected plant host. Where the relationship is not symbiotic, but microbial association with the host is required (e.g. PGPR), specific colonization of the rhizosphere may occur. Law and Lewis (1983) suggested that the internal partner in a mutualistic symbiosis, the inhabitant (e.g. root nodule bacteria), experiences a relatively constant environment within tissues or cells of its host, the exhabitant (e.g. legume). Consequently, they argued that there is no pressure for the inhabitant to change genetically through sexual reproduction. A corollary to their hypothesis is that inhabitants, once symbiotically engaged, will not develop a specific affinity for a particular host genotype. Harley and Smith (1983) and Holl (1983) concluded that there is no selective pressure for either symbiotic partner to evolve a high degree of infection specificity. The obvious parallel which has been used in the development of these arguments is the gene-for-gene theory, originally expounded by plant pathologists (Flor, 1955; Person, 1959; Vanderplank, 1978). The theory is based on the premise that resistance to disease-causing micro-organisms in the plant population is oligogenic, i.e. determined by one or a few genes, and predicts that for each gene determining resistance in the host, there is a specific and related gene determining virulence in the pathogen. In this game of genetic "cat and mouse", there is pressure for the pathogen to mutate to virulent genotypes and for the host to evolve resistant genotypes. Susceptibility of the host to infection by microbes is considered to be a primitive trait and should be a common feature of

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plant populations. Because the mutualistic association is advantageous to both host and microbe, generalists (plants and microbes) able to infect a number of host genotypes (or to be infected by several microbial genotypes) should maintain a selective advantage and, in theory, infection specificity between mutualistic symbionts should not develop. However, several examples of infection specificity between plants and mutualistic micro-organisms are described in the literature (see following pages), and the reason for its occurrence may be related to the effectiveness of specific plant-micro-organism combinations. In a symbiosis, the microbial inhabitant is not under pressure to reorganize genes sexually (Law and Lewis, 1983); nevertheless, random mutation may be a significant source of variability in such microbial populations. As a general rule, any one gene can be expected to mutate at cell division with a frequency of about lo-’ (Stanier et a l . , 1976). Given that the rhizosphere population of bacteria, for example, can reach 3 x 10” bacterial cells g-’ of soil (Rouatt and Katznelson, 1961), several mutations per bacterial generation may occur. Most of these mutations will be deleterious and lethal, but the potential for a random beneficial mutation to occur does exist. Under suitable conditions, the generation time of soil bacteria may be measured in hours, rather than weeks or months. Therefore, a superior bacterial genotype may rapidly outcomPete an inferior “parental” genotype and become numerically predominant. It is conceivable that once effective plant-microbe mutualisms have developed, selective pressure to maintain these combinations facilitates the evolution of infection specificity between the mutualists. Furthermore, infection specificity may be much more common than is currently thought. While preferential infection is often not detected under laboratory or glasshouse conditions, using a limited number of experimental organisms, it may occur in nature, when plant hosts are exposed to a more diverse microbial population in their natural environment. This phenomenon has been called “ecological specificity” (Harley and Smith, 1983). and there is evidence for its occurrence within VAM and ectomycorrhizal fungi, and root nodule bacteria (Molina and Trappe, 1982a; Chanway et ul., 1989a; McGonigle and Fitter, 1990). Therefore, the current dogma that infection specificity is not particularly close in many plant-micro-organism mutualisms may result from our inability to impose realistic environmental conditions in experiments designed to study this phenomenon. Once infection (or colonization) of the root (or rhizosphere) occurs, those micro-organisms most effective at promoting their own fitness and that of their host should have a selective advantage. The potentially short generation time characteristic of soil micro-organisms would facilitate rapid proliferation of plant-beneficial microbial genotypes in the

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rhizosphere and the development of effective, and hence specific plantmicrobe combinations. The occurrence of specificity at both stages of the association (infection and effect on plant growth) and the associated ecological consequences will be discussed in this review in relation to a number of beneficial symbiotic and associative plant-micro-organism partnerships.

111. RHIZOBIUM Most members of the Leguminosae (Fabaceae) form root nodules which are capable of fixing atmospheric nitrogen when infected with appropriate Rhizobium or Brudyrhizobium strains. Two important features of this symbiosis are the infectivity of root nodule bacteria, the capacity to form root nodules, and their subsequent effectiveness, or the ability to fix nitrogen once nodules have been formed. Specificity between plant hosts and root nodule bacteria may occur during either nodulation (infectivity) o r nitrogen fixation (effectiveness).

A. Infectivity Specificity between species of Rhizobium and members of the Leguminosae is well known and has resulted in the definition of “cross-inoculation” groups-plant species which are nodulated by one species or strain of Rhizobium but not others (Table 4). The concept of “cross-inoculation” groups is based solely on infectivity but the boundaries between such groups are not discrete. The classification system has been re-

Table 4 Cross-inoculation groups of legume root nodule bacteria Genus

Species

Biov a r 0

Examples of host plant genera

Rhizobium

leguminosarum

viceae trifolii phaseoli

meliloti loti

Pisum, Vicia Trifolium Phaseolirs Medicago, Melilotirs Lotus

japonicum miscellaneoush

Glycine Lupinus, Vigna

Bradyrhizobium

Strains which are biochemically similar but with a different host range. Strains which have not been classified into species or biovars.

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organized to include two genera, Bradyrhizobium and Rhizobium (Jordan, 1982). Genetic determinants of host specificity are frequently carried on plasmids, which are small pieces of DNA in the bacterial cytoplasm that replicate autonomously and may move freely from cell to cell. Further readjustmant of the classification system may therefore be necessary because of plasmid mobility within bacterial populations (Young and Johnston, 1989). The formation of a root nodule by a Rhizobium strain depends on several factors including the influence of the host plant, the strain of Rhizobium , and the various environmental interactions that influence them (Date and Brockwell, 1978). The latter group of factors can include biotic interactions such as competition, antibiosis, and bacteriocin or phage production by other soil microbes including competing strains of Rhizobiurn (Trinick, 1982). Edaphic factors are also important in saprophytic competition with, and survival of, Rhizobium strains. These include soil texture, structure, moisture, temperature, pH, organic matter content, and nutrient levels (Bushby, 1982). However, the broad specificity observed between legumes and Rhizobium of various cross-inoculation groups and the much more defined legume genotypeRhizobium isolate specificity demonstrated within different cross-inoculation groups (Marques Pinto et al., 1974; Mytton and de Felice, 1977) suggests that biotic interactions between the host and microsymbiont are also very important in strain selection. Marques Pinto et al. (1974) compared the ratio of Rhizobium strains in solutions used to inoculate plants with the ratio of strains found colonizing the root surface and the subsequent effect on nodulation frequency of the strains. Several different species of Trifolium and Medicago were tested. Generally, they found that the ratio of strains colonizing the root surface more closely predicted the identity of the strain forming the nodule than did the ratio of strains in the initial inoculum. The ratio of strains in the inoculum was not closely related to the ratio found colonizing the root surface, presumably because strains differed in their ability to colonize root surfaces. However, in some experiments neither the ratio of strains in the inoculum nor colonizing the root surface were correlated with nodule formation, and this has been confirmed in other studies (Means et a l . , 1961; Robinson, 1969; Skredleta and Karimova, 1969; Labandera and Vincent, 1975; Franco and Vincent, 1976). These observations imply that root colonization per se is not the only factor determining nodulation success and that other factors must be involved. This contention is supported by Hardarson and Jones (1979) who demonstrated that the strain preference of parental Trifolium repens L. (white clover) genotypes can be inherited by progeny of the host plants.

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If the biotic component of plant-microbe interactions in the rhizosphere is important in infection specificity, then some type of molecular signal(s) between plant roots and bacteria must exist for specificity to occur. Several authors have suggested that the secretion of root exudates such as homoserine, biotin, or thiamine may selectively stimulate some strains of Rhizobium in the rhizosphere (van Egeraat, 1975; Date and Brockwell, 1978). While such relationships do exist, they appear to lack the level of specificity required for the observed interactions between Rhizobiurn strains and their specific legume host symbionts. More recently, molecular genetic analysis of the legume- Rhizobium system has shown that specific plant signals are released which activate the nodulation genes in the appropriate Rhizobium. Long (1989) has reviewed the convincing evidence that secretion of flavonoids and isoflavonoids from the roots mediates the initial nodulation gene response in the infecting bacteria. These studies have confirmed the importance of plant-microbe communication in early mutual recognition of host and microsymbiont. Physical attachment of Rhizobium cells to root hairs precedes root hair curling, development of the infection thread, and subsequent nodule formation (Dart, 1977). Host specificity is expressed at about the same time as attachment, before penetration of the root hair cell wall (Li and Hubbell, 1969) or the formation of the infection thread (Napoli and Hubbell, 1975). The discovery that in vitro binding of Glycine max L. (Merr) (soybean) seed lectin to Rhizobium was correlated with the ability of these strains to infect soybean roots lead to the lectin hypothesis to explain host-microsymbiont specificity (Hamblin and Kent, 1973; Bohlool and Schmidt, 1974). Lectins are carbohydrate-binding glycoproteins. This hypothesis states that recognition at infection sites involves the binding of highly specific plant lectins to unique carbohydrates found only on the bacterial cell surface (Dazzo and Hubbell, 1982). The reaction envisaged is similar to that occurring between a homologous antibody-antigen combination where the lectin molecule would act as a bridge linking the bacterial cell and plant root hair surfaces (Figure 1). Dazzo and Hubbell (1975) correlated in vitro lectin binding and strain specificity in the Trifofium-Rhizobiutn symbiosis. Binding in these studies was shown to be inhibited by the hapten 2-deoxyglucose. Haptens are low molecular weight molecules which bind to reactive sites on the antibody thereby blocking further immunological reaction of the molecules. The seed lectin, trifoliin, which was originally used to demonstrate specificity, has been shown to be present on root hair surfaces where attachment of the bacterial cell occurs before infection (Dazzo et a f . , 1978). Furthermore, Diaz et a f . (1989) have shown that

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,Lectin

molecules

Root' hair wall

Fig. 1. Hypothesized role of lectin in binding Rhizobiurn to root hairs.

transformation of T . repens root cultures with the pea seed lectin gene allowed nodulation by Rhizohium leguminosarum biovar viceae, normally a symbiont of peas (Table 4). Factors such as pre-incubation of bacteria with root exudates (Bhuvaneswari and Bauer, 1978), the nitrate or ammonium ion concentration in the growth medium, or the age of the bacterial culture used for inoculation may affect lectin binding between plant roots and bacterial cells. For example, increasing the concentration of either NO; or NH,' results in a parallel decrease in levels of detectable trifoliin and the attachment of Rhizohium to roots hairs (Dazzo and Brill, 1978). Hrabak er a l . (1981) found that quiescent cultures of Rhizobium are more likely to bind specifically to root hairs than are actively growing ones. However, there are several reports demonstrating that attachment of Rhizobium to various surfaces, including legume roots, is non-specific and does not involve host lectins (Law et a l . , 1982; Badenoch-Jones et a l . . 1984; Mills and Bauer, 1985; Vesper and Bauer, 1986). The variable and transient nature of lectin binding activity may explain in part the conflicting evidence for the involvement of these glycoproteins in infection specificity. Dazzo and Hubbell (1982) suggest that the brief appearance of lectin binding activity by Bradyrhizohium japonicum between lag and exponential phases of growth demonstrated by Bhuvaneswari et a l . (1977) could be ecologically significant. They suggest that the physiological state of the bacteria in late lag phase may resemble that of quiescent Rhizohium cells first encountering the nutrient-rich legume rhizosphere. Turning on a specific attachment mechanism at that point in the bacterial life cycle would help to ensure that the infection process was initiated. It appears, therefore, that plant lectins make some contribution to host/microsymbiont specificity. Whether or not this phenomenon

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elicits a major response in the specific binding of cells to root hairs requires further investigation. The existence of selection pressures which result in infection specificity between legumes and root nodule bacteria is debatable, but the phenomenon does exist and may be important in determining the distributional range of certain plant species. There are several theories regarding the factors influencing infection specificity and the biochemical mechanisms involved, none of which has been unequivocally substantiated. However, the discovery that genetic determinants of specificity in the root nodule bacteria are carried on exchangeable segments of DNA (plasmids) may lead to a complete re-evaluation of the ecology and taxonomy of these symbiotic micro-organisms.

B. Effectiveness Symbiotic effectiveness is the capacity to fix nitrogen after root nodules have been formed. Host specificity has also been demonstrated within cross-inoculation groups when nitrogen fixation or plant performance has been measured. The occurrence of specificity at this level is not surprising when the complexity and intrinsic variability of plant and bacterial traits influencing the symbiosis is considered. For example, heterogeneity between T. repens genotypes has been demonstrated for a range of biochemical, morphological, and phenological characteristics, including nitrogen fixation (Mytton, 1975; Burdon, 1980). Variability in morphological and physiological traits such as cell wall antigen composition and response to antibiotics have been demonstrated between isolates of the same species of Rhizobium (Trinick, 1969; Mytton, 1975; Mytton and de Felice, 1977; Hagedorn, 1979; Hagedorn and Caldwell, 1981; Dughri and Bottomley, 1983; Chanway and Holl, 1986) and quantitative variation in nitrogen fixation between specific legume- Rhizobium combinations is also well documented (Mytton, 1975; Mytton and de Felice, 1977; Mytton et al. 1977; Hobbs and Mahon, 1983). The complex interaction between the host and microsymbiont ensures that there are many factors in the symbiosis which could affect nitrogen fixation qualitatively and quantitatively. For example, the enzyme nitrogenase is extremely oxygen-labile but the process of nitrogen fixation requires energy (Mahon, 1982). Substantial aerobic respiration must occur in close proximity to the nitrogen-fixing enzyme for generation of ATP to drive the process. Leghaemoglobin, an oxygen carrying protein, regulates this delicate balance between oxygen toxicity and the aerobic energy requirement by simultaneously providing or sequestering oxygen in the nodule as required. The potential for qualitative and quantitative variability in leghaemoglobin is substantial because the haem portion of

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the molecule is coded for by bacterial genes while the protein originates in the plant (Cutting and Schulman, 1971). Differences in genetic compatibility between host and microsymbiont and in leghaemoglobin activity could cause variability in symbiotic effectiveness, i.e. in the amount of nitrogen fixed. Ammonia, the potentially toxic end product of nitrogen fixation, must be rapidly and efficiently removed from the nodule after its formation. Export efficiency of ammonia from the nodule may also depend on compatibility of the host plant and microsymbiont. Therefore, selection of Rhizobium strains through co-existence with the host genotype should be expected to result in the evolution of metabolically compatible and therefore highly effective genotypic combinations of plants and symbiotic bacteria. Mytton (1975) investigated the possibility that effective legume- Rhizobium combinations may evolve, by collecting eight genotypes of T. repens and isolating Rhizobium from the associated root nodules. When T. repens genotypes were inoculated with Rhizobium strains previously isolated from their root nodules, plant dry weight was an average of 25% higher compared with the performance of T. repens genotypes inoculated with unrelated Rhizobium isolated from other T. repens genotypes. Similarly, Mytton and Livesly (1983) showed that the Rhizobium population-plant variety interaction had the largest effect on T. repens yield when several T. repens varieties and microbial strains were tested at different locations with varying soil types. Other studies with Trifolium and Rhizobium have demonstrated comparable results (Sherwood and Masterson, 1974; Young and Mytton, 1983; Mytton and Hughes, 1984; Chanway et al., 1989a). Microbe-legume specificity has also been demonstrated with other plant and Rhizobium species. For example, the interaction between Rhizobium meliloti and alfalfa (Medicago sativa L.) genotypes has been shown to comprise a significant component of the variability in plant yield (Mytton ef al., 1984). Furthermore, experimentation with previously unrelated R. legiiminosarum biovar viceue strains and field bean (Vicia fabu L.) varieties demonstrated that the specific interaction between microbe and legume genotypes accounted for 74% of the variation in plant dry weight (Mytton er al., 1977). Similar effects have been observed with Rhizohium and peas (Pisum sativum L.) (Hobbs and Mahon, 1982; 1983) and crimson clover (Trifolium incarnatum L.) (Smith et al., 1982). However, the environment plays an important role in these interactions as Dean et al. (1980) found no interaction between genotypes of field bean and R . leguminosarum biovar viceae. Large environmental effects have also been noted in work with T. repens and Rhizobium (Mytton, 1981; Mytton and Hughes, 1984).

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C. The Relationship of Infectivity to Effectiveness The relationship between infectivity and effectiveness has been studied using Rhizobium and various species of Trifoliurn. Apparent selection of Rhizobium strains by T. repens has been observed (Masterson and Sherwood, 1974; Marques Pinto et al., 1974; Russell and Jones, 1975; Mytton and de Felice, 1977; Hardarson and Jones, 1979; Jones and Hardarson, 1979) and by Trifolium pratense (Robinson, 1969; Russell and Jones, 1975; Materon and Hagedorn, 1983). In some cases, the most infective Rhizobiurn strains have resulted in the most effective nitrogen fixing associations (Robinson, 1969; Masterson and Sherwood, 1974; Marques Pinto et al., 1974; Materon and Hagedorn, 1983). However, less effective or completely ineffective strains have also been shown sometimes to form the majority of nodules (Vincent and Waters, 1953; Mytton and de Felice, 1977). Chanway et a l . (1989a) found that R. leguminosarum biovar rrifolii formed the fewest nodules when T. repens clones were presented with a three strain mixture of root nodule bacteria including one strain isolated previously from the “parental” Trifoliurn host genotype. Therefore, infectivity and effectiveness may be correlated under certain conditions but host selection of the most effective Rhizobiurn isolate from a mixture of strains is not assured. An argument can be made for the existence of selection pressure favouring the development of specificity between legumes and root nodule bacteria when symbiotic effectiveness is considered as a factor. If the symbiosis enhances the fitness of both plant and bacterial genotypes, then highly specific plant-microbe combinations might be expected to evolve. However, the existence of such selection pressure can be questioned on the basis that, once differentiated into root nodule bacteroids, Rhizobium may not survive nodule senescence (Almon, 1933). A consensus regarding the viability of Rhizobium bacteroids has not been reached, possibly because bacteroids associated with certain legume hosts appear to remain viable, while those associated with other host species do not (Gresshoff et al., 1977; Tsien et al., 1977; Paau et a [ . , 1980). However, the issue is not critical to the argument for selection pressure in the Rhizobium-legume symbiosis. Even if bacteroids do not survive nodule senescence, there is a substantial number of vegetative Rhizobiurn cells (of the same strain as that which formed the nodule) in infection threads and disorganized plant cells that will be capable of reproduction and participation in subsequent evolutionary processes (Paau et al., 1980). Whether those symbioses in which bacteroids remain viable evolve specificity more rapidly than others is a matter of speculation. In this regard, it is interesting to note that in the

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Trifolium-Rhizobium symbiosis, which is thought to retain viable bacteroids (Gresshoff et al., 1977), specificity at the genotype level has been detected by several researchers (Mytton et al., 1977; Chanway et al., 1989a). Results from studies using plant and bacterial genotypes that have entered into symbiosis naturally support the hypothesis that effective Rhizobiurn-legume combinations evolve (Mytton, 1975; Turkington rt a f ., 1988; Chanway et al., 1989a). However, infectivity and effectiveness do not necessarily appear to be related (Vincent and Waters, 1953; Mytton and de Felice, 1977). and the host plant may discriminate against the most effective strains even if organisms share a history of co-existence (Chanway et al., 1989a). These observations may in part be the result of the artificial environmental conditions under which such experiments were performed. Further experimentation under representative ecological conditions is required to assess the fitness of host plant and associated root nodule bacteria when in symbiosis to explain the sometimes anomalous relationship between infectivity and effectiveness, and the role of co-existence.

.

IV FRANKIA Some non-leguminous plants form root nodules which are capable of fixing atmospheric nitrogen. Bradyrhizobium is capable of infecting the non-legume, Parasponia andersonii Planch., a member of the elm family, and nitrogen-fixing cyanobacteria (blue-green algae) nodulate some species in the order Cycadales. However, the micro-organism most often responsible for nitrogen fixation in non-legumes is an actinomycete (filamentous bacterium) belonging to the genus Frankia. A total of 170 plant species belonging to eight families consisting mostly of perennial woody shrubs and trees are known to enter into actinorhizal (i.e. actinomycete) symbiosis (Becking, 1982). Root nodules are analogous to those found on legumes in that they are located in the root system and contain bacteria that symbiotically fix nitrogen. However, there are several important structural and biochemical differences between legume and non-legume root nodules (Table 5 ) . Some Frankia isolates sporulate readily in root nodules and this characteristic has been used to distinguish between two types of actinorhizal root nodules designated sp( +) for those containing numerous spores and sp(-) for nodules containing relatively few spores (van Dijk and Merkus, 1976; van Dijk, 1978). There is evidence that this characteristic is controlled by the Frankia genotype (van Dijk, 1978; Normand and Lalonde, 1982; VandenBosch and Torrey, 1985). However, host

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Table 5 Some differences between legume and non-legume root nodules Characteristic

Legume nodules

Non-legume nodules

Bacterial symbiont Location of bacteria

Rhizobium/Bradyrliizobium

Frankia External to the vascular system Perennial Absent"

~

Nodule phenology Leghaemoglobin

Internal to the root vascular system Generally annual Present

Haemoglobin has been detected in some nodules (Tjepkema. 1983)

and/or environmental influences are also important because most sp( -) strains readily sporulate when cultured in vitro (Burggraaf et al., 1981; Normand and Lalonde, 1982). The ecological significance of sp(+) and sp(-) nodules has not yet been determined but the relative abundance of sp(+) and sp(-) nodules also depends on the environment in which host plants grow (Weber, 1986; Holman and Schwintzer, 1987; Kashanski and Schwintzer, 1987).

A. Infectivity "Cross-inoculation" groups appear to exist within the genus Frankia, but much of the early work was done using crushed nodule suspensions as inocula (Quispel and Burggraaf, 1981). This technique may yield misleading results because nodule suspensions may consist of multiple Frankia strains and/or non- Frankia microbial contaminants which affect nodulation (Postgate, 1982; Richards, 1987). Callaham et al. (1978) were the first to obtain a pure culture isolate of the endophyte and subsequently several methods have been developed to facilitate its isolation (Baker et al., 1979; Quispel and Burggraaf, 1981; Lalonde et al., 1981). Baker (1987) used pure cultures of Frankia isolated from different actinorhizal plants to test for host specific infectivity in a controlled environment. Four host infectivity groups were identified: strains which nodulate Alnus (alder) and Myrica (myrtle); those nodulating Casuarina (ironwood) and Myrica; those nodulating the Elaeagnaceae (Elaeagnus and Hippophue) and Myrica; and strains which nodulate only the Elaeagnaceae. No infection specificity was observed within the genus Alnus when A . glutinosa (L.) Gaertn. and A . rubra Bong. were tested. The lack of infection specificity between Frankia strains and Alnus species has been confirmed in other studies (Weber et a [ . , 1987; Prat, 1989; Steele et a l . , 1989). However, there is some evidence for infection

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specificity by Frankia strains isolated from the Casuarinaceae (Allocasuarina, Casuarina, and Gymnostoma) (Nazaret et al., 1989; Torrey and Racette, 1989). Generally, we know very little about infection specificity associated with the actinorhizal symbioses. It appears that some specificity does occur but whether the genetic determinants are plasmid-borne as is the case with Rhizobium has not been determined. Richards (1987) concluded that “few generalizations about specificity are possible: at most it could be said that there appears to be nothing like the marked specificity found in the legumes.”

B. Effectiveness Differences between Frankia-host combinations have been detected when symbiotic effectiveness measured via acetylene reduction or plant growth has been monitored. Dillon and Baker (1982) found significant differences in acetylene reduction rates when five Frankia strains, previously isolated from Alnus viridis ssp. crispa (Aiton) Turrill, A . viridis ssp. sinuata, Alnus rubra Bong, Comptonia peregrina (L.) Coult., and Myrica pennsylvanica Loisel., respectively, were tested in a factorial experiment with host plants of the same species. In general, differences in acetylene reduction rates were not correlated with the origin of the bacterial isolate, e.g. Frankia originally isolated from A . viridis ssp. crispa did not result in the highest acetylene reduction rate when tested with A . viridis ssp. crispa as host. The single exception was Myrica gale L.: when inoculated with the endophyte isolated from M . pennsylvanica, the mean rate of acetylene reduction was highest. In contrast, Dawson and Sun (1981), Weber et al. (1987), and Prat (1989) found no significant interaction between endophyte source and Alnus host when plant growth was monitored. Similarly, Rosbrook and Bowen (1987) detected no interaction when three Frankia isolates were tested with four species of Casuarinl?. Normand and Lalonde (1982) studied infection and symbiotic specificity of more than 200 Frankia strains isolated from 27 provenances of A . crispa (Ait.) Pursh. and Alnus rugosa (Du Roy) Spreng. Some differences between provenances were found but the major contributing factor to yield differences was the sp(+) or sp(-) endophytic character. In a further study, Simon et al. (1985) found no evidence of a plant clone- Frankia isolate interaction when symbiotic effectiveness of plant-microbe combinations was monitored within the species Alnus glutinosa. There is some evidence that greater specificity may exist between plant hosts and sp(+) isolates of Frankia. Weber et al. (1987) were unable to culture sp(+) Frankia from nodules of Alnus incana L. and

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A . glutinosa, but using crushed nodule suspensions, these authors demonstrated that the source of the sp( +) endophyte affected symbiotic effectiveness. For example, effective nodules were formed on A . glutinosa only when inoculum originated from that crushed nodules of that host. When A . glutinosa was inoculated with crushed nodule suspensions from sp( +) A . incana nodules, only small ineffective pre-nodule structures formed. Similarly, ineffective nodules resulted on C. peregrina when inoculated with a nodule suspension from M . gale (VandenBosch and Torrey, 1983). Specificity associated with the actinorhizal symbiosis is just beginning to be characterized. However, from the limited data base available, it appears that specificity in relation to effectiveness is not a large component of the variability in actinorhizal nitrogen fixation and subsequent plant growth. However, a greater degree of specificity may exist between hosts and sp(+) isolates of Frankia compared with sp(-) isolates. To confirm this proposition, cultural requirements of sp( +) Frankia must be determined to facilitate pure culture experimentation with these isolates.

V. ASSOCIATIVE RHIZOSPHERE BACTERIA A. Infectivity The concept of infectivity is not strictly applicable to a discussion of rhizosphere bacteria that colonize, but do not infect, root systems; a more appropriate term is rhizosphere colonization. There is evidence that such bacteria may “infect” the root cortical tissue and proliferate between cells of the root cortex (Umali-Garcia et al., 1978; Patriquin, 1982). Associative rhizosphere bacteria may also occur within epidermal cells (Lindberg et al., 1985), but these micro-organisms are generally restricted to the mucigel and adhering soil around the roots (the rhizosheath and rhizosphere) and the root surface (the rhizoplane) of host plants (Burris et al., 1978; Larson and Neal, 1978; Wullstein et a[., 1979; Umali-Garcia et al., 1980; O’Hara er a l . , 1983; Lindberg et a [ . , 1985). See Table 1 for definitions of these terms. Rhizosphere colonization is an obvious pre-requisite for PGPR (Table 1) to affect host plant performance (Schmidt, 1979) and many species of soil bacteria are commonly found in the rhizosphere (including the rhizoplane) (Juhnke et al., 1987). The ability of bacteria to colonize the rhizosphere is a characteristic of the bacterial isolate and varies within species (Kloepper et al., 1985). Differential colonization between plants by beneficial soil bacteria has been documented and in some cases the

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source of specificity is attributable to the presence of a single chromosome in the host plant (Neal et a l . , 1970, 1973; Atkinson er a l . , 1974). While intuitively appealing, consistent demonstration of a quantitative relationship between rhizosphere colonization by PGPR and subsequent plant growth promotion is lacking. For example, Suslow and Schroth (1982) concluded that the establishment of high populations of PGPR on roots of sugar beet appeared to be related to plant growth promotion. Bashan (1986a,b) reached a similar conclusion when wheat was inoculated with Azospirillum brasilense. However, Reddy and Rahe (1989) found that Bacillus subrilis strain B-2 failed to maintain a large rhizosphere population after inoculation of onion but caused shoot and root dry weight increases of up to 100%. Recent work with maize PGPR confirmed that the rhizosphere population of growth-promoting Azospirillum is small relative to the total heterotrophic bacterial population associated with maize roots (Mubyana, 1990). It is likely that a threshold rhizosphere population exists beyond which beneficial effects on plant growth are manifested. There is no a priori reason why the threshold population must be larger than that of other rhizosphere micro-organisms; it only need be large enough to elicit the growth response in the plant host.

B. Effectiveness Some strains of PGPR are capable of promoting the growth of a number of plant species (Holl et a l . , 1988; Bashan et a l . , 1989). However, PGPR are not universally effective and differences in growth promotion between PGPR-plant combinations are well documented (Rovira, 1963; Chanway er al., 1989b). The basis of these differences is not understood, but host plant genetics (Chanway et al., 1989b) and the history of co-existence between plant and microbial genotypes (Chanway et al. , 1988a,b) are important. If plants and beneficial rhizosphere bacteria (or bacterial populations) co-evolve, then the fitness of both partners must be greater when growing in association compared with their respective fitnesses when growing alone. Therefore, root-associated micro-organisms (e.g. rhizobacteria) that contribute to the fitness of their host may have a selective advantage over other rhizosphere micro-organisms, and should comprise a significant, but not necessarily dominant, proportion of the rhizosphere microflora. However, the evolutionary relationship of PGPR with plant “hosts” has not generally been considered to be an important determinant of plant growth promotion, although anecdotal evidence which supports this idea has been accumulating (Sumner, 1990). The occurrence of specific relationships between strains of associative

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nitrogen fixing bacteria and plant genotypes is well known (see Holl, 1983). Recently, Chanway et al. (1988a) tested the hypothesis that plant genotypes and associative growth-promoting bacteria (or bacterial populations) may co-evolve, resulting in genotype specific plant growth promotion by rhizosphere bacteria. Pairs of neighbouring T. repens and Lolium perenne plants were collected from a pasture which had been sown 45 years previously and was managed solely by adjustment of the timing and intensity of grazing. Rhizosphere bacteria (Rhizobium and Bacillus) were isolated and tested for their effects on plant growth. Inoculation with Bacillus stimulated Trifolium yield, the magnitude of which depended on the genotypic identity of grass or bacterial neighbours. The term “co-existent” has been used to describe plant genotypes ( T . repens and L. perenne) that were physically contacting neighbours in the field. Bacteria isolated from T. repens root nodules or the associated rhizosphere soil of these plant pairs are considered to be co-existent with genotypes of each of these plant species. The term “unrelated” describes plant or bacterial genotypes that were not co-existent neighbours in the field. As the experimental biotic environment became more “familiar” by growing Trifolium with (i) unrelated Bacillus and Lolium , (ii) co-existent Bacillus. but not Lolium, and (iii) co-existent Bacillus and Lolium (i.e. organisms that had co-existed with the Trijolium genotype in the field), the yield of the legume component of the species mixtures was shown to increase i.e from condition (i) to condition (iii). In contrast, no inoculation response was detected when plant genotypes were inoculated with unrelated Bacillus. Rennie and Larson (1979) demonstrated that the plant response to growth-promoting soil bacteria was dependent on the host plant genotype. They used disomic chromosome substitution lines of spring wheat and two bacterial species, a nitrogen-fixing Bacillus isolate (C-11-25) and A . brasilense sp. 7 (ATCC 29145). (Disomic chromosome substitution lines contain 20 pairs of indigenous chromosomes plus one pair from a donor line. allowing for the study of the effects of the “donated” pair in an otherwise constant genetic background). A highly specific response to bacterial species was observed after 5 weeks of plant growth. For example, inoculation of C-R5D (‘Cadet’ genomic complement with chromosome 5D substitution from ‘Rescue’) with Bacillus isolate C-11-25 resulted in a 59% increase in plant dry weight while the parental cultivar *Cadet’ showed a 28% increase in dry matter due to inoculation. A more striking difference in growth response was observed when the other parental cultivar .Rescue’ and associated substitution lines were tested. Addition of Bacillus isolate C-11-25 to ‘Rescue’ resulted in a 37% reduction in plant dry weight, but when R-C5D was

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tested, inoculation with Bacillus almost tripled plant dry weight. Therefore, relatively small changes in plant genotype can have substantial effects on growth promotion by Bacillus. Similar chromosome-dependent specificity was observed with A . bradense. Johnson and Mattern (1977) and Klucas and Pedersen (1980) found that specific chromosomes of ‘Cheyenne’ and ‘Atlas’ wheat were important in establishment of effective nitrogen-fixing associations. The discovery, that the genotypic match between plants and microbes can determine the nature of the plant growth response, is significant from ecological, and possibly evolutionary, perspectives. Some authors have noted the potential importance of plant genotype effects in growth promotion by bacteria (Millet ef al.. 1985; Sumner, 1990) but until recently, no studies had focussed on the possibility that plant genotypic differences might result in differential selective pressure on rhizosphere populations of beneficial associative bacteria. If benefical micro-organisms are positively selected over time in the plant rhizosphere, then the probability of finding a positive effect on plant performance due to inoculation with co-existent bacterial strains should be greater than if plants are inoculated with strains to which they have not been previously exposed. To test this hypothesis, Chanway ef al. (1988b) chose spring wheat growing on the Canadian prairies for study because fields which have been cultivated solely to this species for several decades were available for experimentation. Several isolates belonging to the genus Bacillus were isolated from the rhizosphere of spring wheat cultivar Katepwa growing in a field which had been continuously cropped to wheat for 27 years. Six of seven isolates tested promoted root growth of ‘Katepwa‘, but not of ‘Neepawa’ or ’HY320’ (Chanway ef al., 1988b). It was originally postulated that some of the isolates might promote growth of the “parental“ ‘Katepwa’ and/or ‘Neepawa’ because the latter cultivar is a close genetic relative of ‘Katepwa’ and was grown in the field from which plant collections were made during years 11-22 of the 27 years that wheat had been grown. These results indicated, however. that specific adaptation of rhizosphere bacteria (or the bacterial population) to wheat can be cultivar-specific and can occur within a period of 5 years. Rennie and Larson (1979) isolated Bacillus strain C-11-25 from soil that had been used to grow wheat since 1911. The Bacillus strains used by Chanway ef al. (1988a) were isolated from a permanent pasture sown in 1939 and managed solely by adjustment of the timing and intensity of grazing, and those used by Chanway et a l . (1988b) were isolated from soil continuously cropped to spring wheat for 27 years and to the same cultivar for the 5 years immediately preceeding sampling. In these three studies, inoculation of plants with Bacillus isolates that had previously

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co-existed with the target plant specifically promoted growth of the co-existent genotype o r variety. There are two possible mechanisms by which genotype specificity between beneficial rhizosphere bacteria and plants may develop. Particular bacterial strains already present in the soil may experience a competitive advantage in the rhizosphere. Consequently, bacterial fitness would increase and the strain would comprise a significant, or possibly dominant, component of the rhizosphere microflora. This would result in the specific adaptation of the soil bacterial population to the rhizosphere of the host plant. A less likely, but possible, mechanism would involve adaptation of particular bacterial strains to the host plant. There are several ways by which bacteria may undergo genetic reorganization (Stanier et al., 1976). These include direct transfer of DNA between cells (i.e. conjugation), transfer of DNA fragments between cells mediated by bacterial viruses (i.e. transduction), and release of DNA fragments from donor cells into growth medium which are subsequently taken up by recipient cells (i.e. transformation) (see Table 1 for definitions). In addition, the process of random mutation will generate several genetic variants per bacterial generation (see Section 11). Therefore, in this scenario, preexisting genetic variability in the bacterial population is not as important as recently generated variability, and selection within the rhizosphere increases the frequency of superior bacterial genotypes. At present, it is not known which of these mechanisms operate in the rhizosphere or if it is some combination of the two. It would be interesting to determine if beneficial bacteria specific to a plant host could be "generated" by repeated culturing of a bacterial strain in the rhizosphere of a genetically stable plant host. Regardless of the mechanism involved, these findings indicate that genotype specificity can be important in determining plant responses to microbial inoculation and that beneficial bacteria or bacterial populations may adapt to specific plant genotypes. More research is needed to determine whether members of the genus Bacillus are unique in their ability to enter into genotype-specific relationships with plants or if this is a more general phenomenon applicable to many, or even most taxa of rhizosphere bacteria.

VI. ECTOMYCORRHIZAL FUNGI Most woody plants form symbiotic partnerships with mycorrhizal fungi (Harley and Smith, 1983: Newman and Reddell, 1987). Benefits accruing to the host plant are numerous and may include (i) increased

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absorption of water and nutrients from soil, (ii) protection from pathogens, (iii) increased tolerance to soil toxins such as heavy metals, and (iv) lengthening of root life (Trappe, 1977; Perry er a l . , 1987). Woody plants, including members of the family Pinaceae, are most commonly infected by ectomycorrhizal fungi (Meyer, 1973).

A. Infectivity Generally, infection specificity is not considered to be close in the mycorrhizal association (Harley and Smith, 1983). However, fungal symbionts vary widely in their ability to enter into symbiosis with conifer species, For example, fungi such as Amanira mriscaria, Boletus ediilis , and Laccaria laccata associate with a diverse range of hosts and have been termed “broad host-range fungi” (Trappe, 1962), while species of Suillus and Rhizopogon have been considered ”host-specific fungi”. Molina and Trappe (1982a) described sporocarp-host association and ectomycorrhiza formation by 27 fungal isolates with seven conifer species from the Pacific Northwest of the USA. Three groups of fungi were described: (i) fungi with wide ectomycorrhizal and sporocarp host potential, (ii) fungi with intermediate host potential but specific sporocarp-host association, and (iii) fungi with narrow ectomycorrhizal and sporocarp-host potential (Table 6). Those belonging to the latter two groups may include “ecologically specific” (sensu Harley and Smith, 1983) mycorrhizal fungi. Variability in specificity of host plants has also been demonstrated. The ericaceous hosts Arbiitiis menziesii Pursh (Pacific madrone) and Arctostaphylos uva-ursi (L.) Spreng. (bearberry) have been shown to form mycorrhizas with a number of fungal associates while Alnrrs species are much more restrictive towards their fungal symbionts (Moh a , 1979, 1981; Molina and Trappe, 1982b). Results from work with different ectomycorrhizal fungi and host plants generally tend to support the idea that there is no selection pressure for infection specificity to evolve between organisms. However, certain trees may support host specific fungi (Malajczuk er al., 1982; Molina and Trappe, 1982a). therefore Richards (1987) concluded that generalizations about the lack of specificity in ectomycorrhizal infections may be be somewhat premature.

B. Effectiveness Differences in the infectivity of ectomycorrhizal fungi with various hosts do not necessarily relate to the effectiveness of the resulting mycorrhiza. Several studies have shown that plant performance (usually monitored

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Table 6 Specificity of some ectomycorrhizal fungi (adapted from Molina and Trappe, 1982a)

Degree of specificity

Fungal species

Conifer host species"

Wide host potential

Amanita muscaria Boletus edulis Cenococcum geophilum Laccaria laccata Paxillus inr3okitus Pisolithus tinctorius

Pme ,Lo ,Th ,Ps ,Pm ,Pp ,Pc Pme ,Lo ,Th ,Ps,Pm ,Pp ,Pc Pme ,Lo ,Th ,Ps,Pm ,Pp ,Pc Pme ,Lo,Th ,Ps,Pm ,Pp ,Pc Pme ,Lo ,Th ,Ps,Pm ,Pp ,Pc Pme ,Lo,Th ,Ps,Pm ,Pp ,Pc

Intermediate host potential

Suillus lakei Sirillus brevipes Rhizopogon vinicolor Rhizopogon occidentalis A lpo va diplophloeus Cortinarius pistorius Rhizopogon cokeri Gastroboletus subalpinus Truncocolumella citrina Zelleromyces gilkeyea

Pme ,Lo ,Th ,Ps,Pp,Pc Pme ,Lo,Ps ,Pm ,Pp ,Pc Pme,Lo,Th,Ps,Pm Ps,Pm,Pc

Narrow host potential

h

Pm,Pp ,Pc Pm,Pp ,Pc Pme,Pm,Pp,Pc Pme ,Lo Pme ,Th

0 Seven conifer hosts were tested: abbreviarions: Pme, Pseridorsttga menziesii: Lo. Larix occidentalis: Th. Tsuga heterophylla: Ps. Picea sirchensis; Pm, Pinus monticola: Pp. Pinus ponderosa; Pc. Pinrts contorta. Forms ectomycorrhizas with AInirs spp. but not with any of the conifers tested.

by dry weight increment of the host) after mycorrhizal formation varies with the species of fungus that has infected the root (Trappe, 1977; Cline and Reid, 1982). However, ecotypic variation with respect to a variety of traits within fungal species has also been documented (Trappe, 1977) leading to the suggestion that the provenance of the fungus as well as that of the tree should be considered in any inoculation programme. It is also well known that mycorrhiza-forming tree species normally support a number of different fungal symbionts at once (Perry et a l . , 1987). However, Malajczuk et al. (1982) found that host-specific fungi may predominate as trees mature, which indicates that certain fungal strains have a selective advantage over others. The failure of southern isolates of Pisolithus tinctorius to improve conifer performance in the Pacific Northwest supports the idea that adaptation of mycorrhizal fungi to the local physical and/or biotic environment is important (Perry et a l . , 1987). Harley and Smith (1983) emphasize that future work on specificity should include an evaluation of host-fungus combinations at the level of fungal strains and host genotypes as well as at higher

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taxonomic levels and that processes such as nutrient uptake per unit root length or carbohydrate consumption may more meaningfully indicate effectiveness rather than the standard plant dry weight measurements.

VII. VESICULAR-ARBUSCULAR MYCORRHIZAL (VAM) FUNGI A. Infectivity Generally, a VAM fungus isolated from one species of host plant can be expected to infect any other species which has been shown to be capable of forming VA mycorrhizas (Mosse, 1978: Hayman, 1982). However, Graw et al. (1979) have suggested that certain fungi have restricted host ranges and differential infectivity studies with plant cultivars o r genotypes and GIomiis species support this idea. Azcon and Ocampo (1981) tested 13 wheat cultivars with a culture of Glornus niosseae and observed a cultivar-dependent range of VAM infection on roots, but two cultivars remained uninfected after inoculation. Lioi and Giovannetti (1987) had similar results when ten ecotypes of Hedysarurii coronariuin L. (sulla) were inoculated with Gloniiis caledonium. Krishna ef ul. (1985) also observed a range in degree of VAM infection when 30 Penni.setuin americanun? (L.) Leeke (pearl millet) genotypes were grown in soils known to contain natural populations of four VAM fungal genera. None of the plant genotypes was completely resistant to infection possibly due to the multiplicity of fungal genera present in the soil. Mercy el a l . , (1990) found that colonization of cowpea (Vigna unguiculata L.) genotypes was host dependent and heritable when plants were grown in soil known to contain 800 infective VAM propagules g-' . In contrast, Estaun et al. (1987) found no genotypic effects when mycorrhizal infection of three pea genotypes was assessed in response to inoculation with three species of Glomus. Recently, McGonigle and Fitter (1990) demonstrated that "ecological specificity" can occur in VA mycorrhizal associations as well. In a field survey of endomycorrhizas in a hay meadow, it was found that Holcus lanatits L. was predominantly infected by Glornus tenue, while neighbouring Ranitnculri.s acris L., Plantago lanceolata L., and Phleurn yratense L. were colonized mostly by other "coarse endophytes", which were physically distinguishable from G. renue. As is the case with the ectomycorrhizal symbiosis, the current view that infection specificity is almost completely lacking in VAM associations may require reassessment.

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B. Effectiveness Differences between fungal strains or species in altering host plant growth are well documented (Mosse, 1972; Bevege and Bowen, 1975; Abbott and Robson, 1984), but detailed investigations of the response of cultivars or plant genotypes to inoculation with VAM are few. When genotypes of pea, wheat, pearl millet, and maize or species of Cifrus and Triticum were studied, VAM fungi have been shown to differ in their symbiotic effectiveness (Hall, 1978; Berthau et a l . , 1980; Menge et al., 1980; Azcon and Ocampo, 1981; Krishna et al., 1985; Estaun ef a l . , 1987). The response of host plants to inoculation includes growth stimulation, reduction, or no detectable effect on growth. Dixon (1988) observed a significant host plant X mycorrhizal fungus interaction in the growth of black walnut (Jugluns nigru L.) when plants originating from three seed sources were inoculated with different VAM species. In contrast, no varietal differences due to VAM inoculation were detected when T. repens or M . safiva were studied (Crush and Caradus, 1980; O'Bannon et al.. 1980). Inoculation techniques and environmental conditions under which experiments are performed have not been standardized. Therefore, it is possible that apparent differences in effectiveness may really reflect differences in the rate and amount of infection or the quantity of propagules in the original inoculum (Abbott and Robson, 1984). Barea and Azcon-Aguilar (1983) concluded that such differences may be related more to the specific soil-plant system they colonize than to the genotype of the host itself. Clearly, more research using standardized techniques is required to understand specificity (or the lack of it) between host and fungus in VAM plants.

VIII. EFFECTS ON PLANT COMPETITION A. Symbiotic Bacteria As previously discussed, natural selection should result in specificity with respect to the effectiveness of mutualistic plant-microbe combinations. We have reported evidence for the existence of such fine-scale specialization which affects plant competitive relationships within L. perenne-T. repens dominated areas of a permanent pasture (Turkington et al., 1988: Chanway et a l . , 1989a). Turkington and Harper (1979) and Aarssen and Turkington (1985) reported that T. repens populations can differentiate into subpopulations, each one specific to the species or genotype of its particular grass neighbour. Analysis of the above-ground behaviour of these plants

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offered no obvious explanation for such compatibility. Many studies have shown that the abundance of rhizosphere and rhizoplane microorganisms, including symbiotic and associative bacteria and mycorrhizal fungi, is affected by neighbouring vegetation (Krasil’nikov, 1958; Rovira, 1961; Tuzimura and Watanabe, 1962; Robinson 1967; Chatel and Greenwood, 1973; Christie et al., 1978; Lawley et al., 1982). Therefore, the mechanism by which genotypic specificity occurs was predicted to involve rhizosphere effects. The work by Mytton and colleagues (Mytton, 1975; Mytton and de Felice, 1977; Mytton et a l . , 1977; Mytton and Livesly, 1983; Young and Mytton, 1983; Mytton and Hughes, 1984) suggests that the relationship between T. repens and Rhizobium might be an important determinant of the genotype specificity documented by Aarssen and Turkington (1985). Therefore, we designed experiments to test the hypothesis that the origin of the root nodule bacteria specifically affects growth of mixtures of genotypes of T. repens and L. perenne (Turkington et a f . , 1988; Chanway et al., 1989a). In an initial set of experiments, it was found that the yield of T. repens was significantly influenced not only by the identity of the Rhizobium isolate infecting root nodules, but also by the presence of the specific combination of Rhizobium strain and grass neighbour (Thompson et al., 1990). A subsequent set of experiments was conducted to investigate the role of Rhizobiurn in the L. perenne neighbour-specific differentiation of T. repens populations. Three pairs of plants collected from widely separated L. perenne-dominated areas of a 45-year-old pasture were used. Each pair was selected as a single ramet of L. perenne and a single ramet of T. repens growing in close proximity. Rhizobium leguminosarum biovar trifolii was isolated from root nodules of each “parental” T. repens plant. This sampling approach resulted in three sets of experimental organisms arbitrarily designated 1. 2, and 3 , each set consisting of a L. perenne plant (L), a strain of Rhizobium (R) isolated from a root nodule of the T. repens (T) plant. Plants were maintained by vegetative propagation in a glasshouse. Combinations of genotypes (plants and bacteria) which were collected from the same microsite in the field were described by the term “co-existent”. A complete factorial experiment was performed in which T. repens growth, as affected by the genotypic identity of the neighbollring L. yerenne plants and Rhizohiirm inoculum, was assessed during a l-year growth period. Little evidence of the genotypic compatibility observed by Aarssen and Turkington (1985) was apparent when the effect of Rhizobium strains was ignored. However, when co-existent Rhizobilrrn was used to inoculate T. perenne genotypes, significant yield advantages were observed regardless of the origin of the T. reyens genotype. The

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effect was manifested initially in T. repens dry weights which were highest when grown with co-existent L. perenne-Rhizobium combinations i.e. LlR1, L2R2, or L3R3. At the end of the experiment, this trend was also detected when dry weight of L. perenne was assessed. Co-existent Rhizobium also conferred a yield advantage to T. repens but the magnitude was substantially less than that seen when co-existent Rhizobium-L. perenne combinations were present (Turkington et a l . , 1988). In another experiment, genotypic mixtures of the pasture plants were inoculated with all three Rhizobium strains to determine if co-existent Rhizobium would form most of the root nodules on T. repens hosts (Chanway et al. 1989a). However, when the influence of T. repens genotypes on Rhizobium strain selection was examined, T. repens plants were found to harbour unrelated microsymbionts in their root nodules in preference to co-existent ones (Fig. 2). The identity of the L. perenne genotypes did not affect this process. The exclusion of co-existent Rhizobium strains did not appear to be of benefit to the host because T. repens yield advantages occurred when co-existent Rhizobium was used as inoculum (Fig. 3; Mytton, 1975). These data support the contention that infectivity and effectivity of Rhizobium strains are not correlated

2.5

1R2 U

0 f

1.5

T1 T 2 T 3

T1 T2 T 3

I

T1 T 2 T 3

Fig. 2. Mean number of Trifolium repens nodules (out of a possible three) infected with co-existent (shaded bars) or unrelated (open bars) strains of Rhizobium leguminosarum biovar trifolii. TI, T2, and T3 are Trifolium genotypes 1, 2, and 3, respectively. R1, R2, and R3 are the corresponding co-existent Rhizobium genotypes 1, 2, and 3 isolated from TI, T2,and T3, respectively. The difference between the number of nodules formed by co-existent Rhizobium (R) compared with the number of nodules formed by unrelated R genotypes is significant ( p < 0.05).

’

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1

1

1

I

I

I

1

I

I

2

3

4

5

6

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0

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Cumulative harvest

Fig. 3. The influence of Rhizobium leguminosarum biovar trifolii strain on the dry weight of Trifolium repens over nine successive harvests over a 1 year period. The yield advantage is a comparative measure of the percentage increased yield of Trifolium when grown with co-existent Rhizobium compared with the mean of all unrelated Rhizobium- Trifolium genotype combinations. Cumulative harvest refers to the total Trifolium biomass accumulated up to each harvest (e.g. cumulative harvest 4 = the total Trifolium biomass accumulated over harvests 1 + 2 + 3 + 4). Solid symbols represent yield advantages significantly greater than zero ( p < 0.05).

(Vincent and Waters, 1953; Mytton and de Felice, 1977) but the ecological significance of the exclusion of co-existent Rhizobiurn strains by host plants can not be ascertained from these data. These results do indicate, however, that genotype communication between plant root and bacterium does occur because co-existent strains were obviously discriminated against in root nodule formation. Results from these experiments suggest that Rhizobium strains are adapted more to L. perenne neighbours than T. repens hosts. Furthermore, plant performance and possibly species distribution in these communities appears to depend on the unexpected and counter-intuitive matching of grass genotype and root nodule bacteria strain which illustrates the complex and unpredictable nature of these neighbour relationships.

B. Associative Rhizosphere Bacteria We have also reported the existence of plant-microbe specificity among members of the genus Bacillus and genotypes of T. repens and L. perenne (Chanway et a l . , 1990). A strain of Bacillus was isolated from

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rhizosphere soil of each of the plant pairs described above and used as inoculum for experiments. Members of the genus Bacillus were selected for the study because nitrogen fixation was hypothesized to be an important component of the phenomenon under investigation and they were the only nitrogen fixers common to all three rhizosphere soil samples. Again, the grass-bacteria interaction was shown to influence above- and below-ground plant performance. Although the results were not as clear as those of the Rhizobium work (Chanway er a l . , 1989a), T. repens growth was best when grown with co-existent L . perenne-Bacillus combinations. This effect was also observed when root nodule weight of T. repens was assessed. Unlike the results obtained in the Rhizobium experiment, L . perenne also grew best when inoculated with co-existent bacteria (Bacillus). We also tested the effects of Bacillus inoculation on nodule occupancy of T. repens (Fig. 4). It was found that Rhizobium isolates which had co-existed in the field with the Bacillus isolate used as inoculum formed most of the root nodules regardless of the identity of the L . perenne or T. repens genotypes. This is in contrast to the repulsion of co-existent strains when T. repens was tested in pure stand without Bacillus (Fig. 2). Therefore, specificity at the genotype level between species of soil bacteria may also be important in these communities. Rhizosphere

R2

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1.4

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3

0.8 0.6

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=

0.4 0.2

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IIl 81 82 8 3

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Fig.4. Mean number of Trifolium repens nodules (out of a possible three) infected by Rhizobium leguminosarum biovar trifolii when co-inoculated with co-existent (shaded bars) and unrelated (open bars) Baciflus (B). The difference between the number of nodules formed by co-existent Rhizobium (R) compared with the number of nodules formed by urrelated R genotypes is not significant ( p > 0.05).

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bacteria clearly influenced plant growth (and may influence fitness) in these experiments. The presence or absence of these micro-organisms may influence plant species distribution in similar environments.

C. Mycorrhizal Fungi Plant competitive relationships are also affected by mycorrhizal fungi. Inoculation of a Douglas fir (Pseudotsuga menziesii (Mirb.) Franco) monoculture with the ectomycorrhizal fungus Rhizopogon vinicolor resulted in a significant decrease in productivity when compared with an uninoculated stand (Perry et a1 ., 1987). When Douglas fir comprised 50% or 75% of a species mixture with Ponderosa pine ( Pinus ponderosa Doug!.), inoculation with R. vinicolor had no effect on Douglas fir performance. However, when Douglas fir comprised only 25% of the species mixture, inoculation with this fungus resulted in a significant biomass yield increase of the Douglas fir component. Ponderosa pine was unaffected by inoculation with this fungal isolate whether grown in monoculture or in mixture with Douglas fir. Therefore, mycorrhizal fungi had no effect on Ponderosa pine, a detrimental effect on Douglas fir when competition was intraspecific, but a beneficial effect on Douglas fir when most competition was interspecific. In subsequent work, soil pasteurization before seedlings were planted was found to have a substantial effect on the outcome of seedlings competition experiments involving mycorrhizal fungi (Perry et a l . , 1989b), which emphasizes further the importance of soil biota in plant interactions. Inoculation of species mixtures with VAM fungi also affects plant competitive relationships. When L. perenne and H. lanutus were grown together in pots, the outcome was influenced by mycorrhizal inoculation (Fitter, 1977). Holcus lanutus outcompeted L. perenne when inoculated because of a 60% reduction in root growth of L . perenne when mycorrhizal. Similarly, Ocampo (1986) demonstrated that VAM-inoculated sorghum competed more effectively with cabbage than did nonmycorrhizal sorghum. These observations parallel those of Turkington et al. (1988) who noted that inoculation with Rhizobium improved the competitive ability of T. repens with the grasses Holcus and Dactylis. Results from these studies illustrate the complexity of the microbial influence on plant competition and the dynamic influence rhizosphere micro-organisms may exert in natural plant communities. We have demonstrated the importance of specificity with symbiotic or associative bacteria and agricultural annuals or perennials (Chanway et a l . , 1988a,b, 1989a,b, 1990). Many examples from soil transfer experiments suggest that any infective mycorrhizal fungus will alleviate growth

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and survival problems of certain plant species in the short term. However, the importance of “fine-tuning” the relationship through selection in the rhizosphere over the long-term is unknown. There is clear evidence that the pattern of mycorrhizas associated with Betula spp. is ordered in time and in space, and that succession from “early stage” to “late stage” symbiotic fungi occurs in these ecosystems (Mason et a f . , 1983; Last et al., 1983, 1984). Furthermore, some trees harbour host-specific fungi as they age (Malajczuk et a l . , 1982), which suggests that selection for effective fungal symbionts also occurs in the rhizosphere. However, the degree, to which the effectiveness of specific plant-micro-organism combinations is important in many of these examples remains unknown.

IX. CONSEQUENCES OF SPECIFICITY ON PLANT COMMUNITY STRUCTURE The role of rhizosphere micro-organisms on plant species distribution and community structure is a neglected subject in plant population biology (Law, 1988) yet certain lines of evidence suggest that these microbes represent the limiting factor in the distributional range of many plant species (see Perry et al., 1989a for a review). Newman (1978) suggested four ways by which micro-organisms could alter the balance between plant species: (i) a micro-organism may favour one species more than another (e.g. specificity between root nodule bacteria or PGPR and host plants), and the effect may be intensified when plants compete; (ii) one plant may affect the micro-organisms of a neighbouring plant whose performance or fitness is specifically affected by its associated microflora; (iii) micro-organisms may detoxify poisonous materials produced by plants; and (iv) mycorrhizal links between plants may result in direct transfer of nutrients between competing species. In scenarios (i) and (ii), specificity in effectiveness of plantmicro-organism combinations is implicit. In (iii) and (iv), it is not necessary but could occur. We have previously discussed the influence of soil micro-organisms on plant competition and concluded that grasses may influence the differentiation of T. repens into subpopulations through an indirect effect on rhizosphere bacteria (Turkington et al., 1988; Chanway et a l . , 1989a, 1990; Thompson et af., 1990). However, the possibility that specificity (with respect to plant growth) may evolve between plants and root-associated micro-organisms in natural communities has generally not been addressed in ecological research.

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Lie et al. (1987) and Young and Johnston (1989) discuss evidence for infection specificity between legumes and root nodule bacteria from an evolutionary viewpoint. Examples are cited which demonstrate the existence of a range of specificities from very wide to very narrow (i.e. genotype-specific). One scenario in which root-associated micro-organisms may affect the distribution of plant species would occur if the appropriate strain of root nodule bacteria (infective and effective) was absent from a nitrogen poor site thereby preventing a nitrogen-fixing host species from colonizing that site. The most dramatic demonstrations of the influence of soil microflora on the distribution of plant species are found in studies of artificially manipulated or disturbed ecosystems. There are several examples of afforested sites requiring soil transfer from forests to facilitate plantation establishment (Mikola, 1970). The identity of the agent(s) added through soil transfer is not always known, but mycorrhizal fungi are usually implicated (Perry et a l . , 1987). Perry and co-workers (Perry et a l . , 1984, 1989a, Perry and Rose, 1983; Amaranthus and Perry, 1987) have studied forest ecosytems in the Klamath Mountain region of northern California and southern Oregon. One clear-cut, at Cedar Camp, was planted with Douglas fir (Pseudorsuga menziesii) four times and each time without success. However, addition of less than 150 ml of soil from the rooting zone of established conifer plantations to planting holes at Cedar Camp increased survival by ca 50% and doubled the growth rate in the first year after outplanting (Amaranthus and Perry, 1989a). Furthermore, only those seedlings receiving soil transfers were alive 3 years after planting. In another study, addition of root-zone soil from Pacific madrone ( A . rnenziesii Pursh) caused a specific growth response in planted Douglas fir that depended on the vegetation type previously found on the planting site (Amaranthus and Perry, 1989b). Conifer seedlings planted on sites previously dominated by Oregon white oak (Quercus garryana Doug]. ex Hook.) or annual meadow species (Cynosurus echinatus L., Galium aparine L . , Poa trivialis L., and Anthriscus sp.) did not respond to the addition of madrone soil. However, addition of madrone soil to conifers planted on sites previously dominated by Arctostaphylos spp. significantly increased seedling growth, the number of mycorrhizal root tips, and the rate of acetylene reduction in the rhizosphere (Amaranthus and Perry, 1989~).The madrone soil influence was thought to be biotic because addition of pasteurized soil did not stimulate Douglas fir growth, mycorrhizal infection, or acetylene reduction. Similar examples have been described when harsh environments such as mine spoils or highly eroded sites have been planted to forest tree seedlings. Without the appropriate fungal partner, seedling growth is

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poor or non-existent (Marx 1975, 1980; Valdes 1986). In a recent review, Perry et a l . (1989a) provide several further examples of changing species composition and degradation of ecosystems due to elimination of important rhizosphere micro-organisms (e.g. mycorrhizal fungi and associative nitrogen fixing bacteria) in response to human-mediated disturbance. Tranquillini (1979) concluded that the range of high elevation conifer species is dependent on ectomycorrhizal fungi adapted to cool temperatures. It appears that without the appropriate mycorrhizal partner, forest seedlings will not grow in certain environments. These studies illustrate the importance of the appropriate rhizosphere microorganisms in determining the range of plant species and in stabilizing floral diversity. Survival, and in some cases proliferation, of microbial mutualists in the rhizosphere of non-host plants may be a very important influence of the distribution of certain plant species. Smolander and Sundman (1987) found high soil populations of Frankia under Betula pendula (Roth) stands known to have been devoid of actinorhizal plants for decades. Subsequent work revealed that Frankia populations were higher under some Betula stands than under stands of the host plant (i.e. Alnus) (Smolander. 1990). A degree of specificity was noted because soils under other tree species (e.g. conifers) had very low populations of these root nodule bacteria (Smolander, 1990). Similar relationships between Rhizobiurn (Robinson 1967; Chatel and Greenwood, 1973; Chanway et al., 1989a) or ectomycorrhizal fungi (Amaranthus and Perry, 1989b) and non-host vegetation have been demonstrated. Connell and Lowman (1989) recently discussed the potential role of mycorrhizal fungi in plant succession and species diversity of tropical rain forests with representative climatic and edaphic conditions. They hypothesize that the relatively close infection specificity and superiority of ectomycorrhizas compared with endomycorrhizas results in low species diversity in some rain forests because the occurrence of dominant species is restricted to those capable of forming ectomycorrhizas with fungal symbionts already present. Experimentation in these ecosystems is still required to test this hypothesis. Plants of the same (Hirrel and Gerdemann, 1979; Whittingham and Read, 1982) or different (Heap and Newman, 1980; Chiariello er al., 1982; Francis and Read, 1984) species can be physically connected by mycorrhizal hyphae. This raises the intriguing possibility that entire plant communities may be physically linked through mycorrhizal connections (see Newman, 1988 for a review). Furthermore, the rhizosphere of mycorrhizal plants (i .e. the mycorrhizosphere) commonly supports several genera of plant benefical bacteria, such as nitrogen fixers or antibiotic producers (Perry et a l . , 1989a), and we have evidence that

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selection within the rhizosphere may result in specific and effective plant-microbe combinations (Chanway et al., 1988a,b, 1989a, 1990). In addition, we have known for some time that biotic (e.g. competition) and abiotic (e.g. soil nutrient status) environmental conditions are very important in determining the growth response of the host plant (Fitter, 1985; Perry et a / . . 1987; Turkington et a l . , 1988). Therefore, plant community structure may in large part result from a fine balance of evolved relationships between plants and associated rhizosphere microflora. Much more research is required to assess the relative influence of rhizosphere micro-organisms (symbiotic and associative) on plant competition, species distribution and community structure in different environments.

X. CONCLUSIONS Results from controlled environment and field studies indicate that the impact of rhizosphere micro-organisms on plant competition and community structure may be substantial, but data are scarce. Arguments based on natural selection have suggested that infection specificity should not develop in mutualistic plant-micro-organism associations (Harley and Smith, 1983). Law and Lewis (1983) conclude that there is selection pressure against continuing evolutionary change on mutualistic inhabitants once symbiotically engaged because of the relatively constant environment experienced within exhabitant tissue or cells. However, infection specificity can readily be demonstrated with root nodule bacteria (e.g. Rhizobiurn) and less so with mycorrhizal fungi (e.g. some species of ectomycorrhizal fungi), and may result from the specific association of effective plant-microbe combinations. Specificity in relation to the effectiveness of mutualistic plant-micro-organism associations would be predicted to evolve if the fitness of both organisms is promoted, and specificity at this level of association has been demonstrated between plant species, cultivars, and genotypes and isolates of rhizosphere bacteria or mycorrhizal fungi. The importance of specificity during infection (infectivity) or subsequent plant growth (effectiveness) on plant competitive ability and subsequent community structure is unknown, but the potential influence is great. A major interdisciplinary research effort involving plant population biology and rhizosphere microbiology is required to further elucidate the role of rhizosphere micro-organisms (symbiotic and associative) and evolved specificity on plant competition, species distribution and community structure in different environments.

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ACKNOWLEDGEMENTS We wish to acknowledge an anonymous referee and A . H. Fitter for thoughtful review of this manuscript and numerous helpful suggestions.

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Perry, D. A., Margolis, H., Choquette, C . , Molina, R. and Trappe, J. M. (1989b). Ectomycorrhizal mediation of competition between coniferous tree species. New Phytol. 112, 501-511. Person, C. (1959). Gene-for-gene relationship’s in host:parasite systems. Can. J . Bot. 37, 1101-1130. Postgate, J . R. (1982). The Fundamentals of Nitrogen Fixation. Cambridge University Press, Cambridge. Prat, D. (1989). Effects of some pure and mixed Frankia strains on seedling growth in different AInus species. Plant Soil 113, 31-38. Quispel, A. and Burggraaf, A. J. P. (1981). Frankia, the diazotrophic endophyte from actinorhizas. In: Current Perspectives in Nitrogen Fixation (Ed. by A. H. Gibson and W. E . Newton), pp. 229- 236. Australian Academy of Sciences, Canberra. Reddy, M. S. and Rahe, J. E. (1989). Growth effects associated with seed bacterization not correlated with populations of Bacillus subtilis inoculant in onion seedling rhizospheres. Soil Biol. Biochem. 21, 373-378. Rennie, R. J . and Larson, R. I. (1979). Nitrogen fixation associated with disomic chromosome substitution lines of spring wheat. Can. J. Bot. 57, 2771-2775. Rennie, R. J. and Thomas, J . B. (1987). lSN-Determined effect of inoculation with N2fixing bacteria on nitrogen assimilation in Western Canadian wheats. Plant Soil 100, 213-223. Richards, B . N . (1987). The Microbiology of Terrestrial Ecosystem. John Wiley, New York. Robinson, A. C. (1967). The influence of host on soil and rhizosphere populations of clover and lucerne nodule bacteria in the field. J . Aust. Inst. Agric. Sci. 33, 207-209. Robinson, A. C. (1969). Host selection for effective Rhizobium trifolii by red clover and subterranean clover in the field. Aust. J. Agr. Res. 20, 1053-1060. Rosbrook, P. A. and Bowen, G. D. (1987). The abilities of three Frankia isolates to nodulate and fix nitrogen with four species of Casuarina. Physiol. Plant. 70, 373-377. Rouatt, J. W. and Katznelson, H. (1961). A study of bacteria on the root surface and in the rhizosphere soil of crop plants. J . Appl. Bact. 24, 164-171. Rovira, A. D. (1961). Rhizobium numbers in the rhizosphere of red clover and Paspulum in relation to soil treatment and the numbers of bacteria and fungi. Aust. J. Agr. Res. 12, 77-83. Rovira, A. D. (1963). Microbial inoculation of plants. 1. Establishment of free-living nitrogen-fixing bacteria in the rhizosphere and their effects on maize, tomato, and wheat. Plant Soil 19, 304-314. Rovira, A. D. (1969). Plant root exudates. Bot. Rev. 35, 35-57. Rovira, A. D. and Davey, C. B. (1974). Biology of the rhizosphere. In: The Plant Root and its Environment (Ed. by E. W. Carson), pp. 153-204. University Press of Virginia, Charlottesville. Russell, E. W. (1973). Soil Conditions and Plant Growth 10th edition. Longmans, London. Russell, P. E. and Jones, D. G . (1975). Variation in the selection of Rhizobium trifolii by varieties of red and white clover. Soil Biol. Biochem. 7, 15-18. Schmidt, E. L. (1979). Initiation of plant root-microbe interactions. Ann. Rev. Microbiol. 33, 355-376. Schroth, M . N. and Weinhold, A. R. (1986). Root-colonizing bacteria and plant health. Hortsci. 21, 1295-1298.

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Sherwood, M. T. and Masterson, C. L. (1974). Importance of using the correct test host in assessing the effectiveness of indigenous populations of Rhizobium trifolii. Ir. J . Agric. Res. 13, 101-108. Simon, L. Stein, A.. Cote, S. and Lalonde, M. (1985). Performance of in vitro propagated Alnus glutinosa (L. ) Gaertn. clones inoculated with Frankiae. Plant Soil 87, 125-133. Skredleta, V. and Karimova, J. (1969). Competition between two somatic serotypes of Rhizobium japonicum used as double-strain inocula in varying proportions. Arch. Mikrobiol. 66, 25-28. Smith, G. R., Knight, W. E., Petersen, H. H. and Hagedorn, C. (1982). The effect of Rhizobium trifolii strains and crimson clover genotypes on N fixation. Crop Sci. 22, 970-972. Smolander, A. (1990). Frankia populations in soils under different tree species with special emphasis on soils under Betula pendula. Plant Soil 121, 1-10. Smolander, A. and Sundman, V. (1987). Frankia in acid soils of forests devoid of actinorhizal plants. Physiol. Plant. 70, 297-303. Stanier, R. Y., Adelberg, E. A . and Ingraham, J . L. (1976). The Microbial World. Prentice Hall, New Jersey. Steele, D. B., Ramirez, K. and Stowers, M. D. (1989). Host plant growth response to inoculation with Frankia. Plant Soil 118, 139-143. Subba Rao, N. S. (1982). Phosphate solubilization by soil microorganisms. In: Advances in Agricultural Microbiology (Ed. by N. S. Subba Rao), pp. 295-303. Butterworth, London. Sumner, M. E. (1990). Crop responses to Azospirillum inoculation. A d v . Soil Sci. 12, 53-123. Suslow, T. V. (1982). Role of root-colonizing bacteria in plant growth. In: Phytopathogenic Prokaryotes Vol. 1 (Ed. by M. S. Mount and G. H. Lacy), pp. 187-223. Academic Press, New York. Suslow, T. V. and Schroth, M. N. (1982). Role of deleterious rhizobacteria as minor pathogens in reducing crop growth. Phytopath. 72, 111-115. Thompson, J. D., Turkington, R. and Holl, F. B. (1990). The influence of Rhizobium leguminosarum biovar trifolii on the growth and neighbour relationships of Trifolium repens and three grasses. Can. J . Bot. 68, 296-303. Tien, T. M., Gaskins, M. H. and Hubbell, D. H. (1979). Plant growth substances produced by Azospirillum brasilense and their effect on the growth of pearl millet (Pennisetum americanum L.). Appl. Env. Microbiol. 37, 1016- 1024. Tjepkema, J. D. (1983). Hemoglobins in the nitrogen-fixing root nodules of actinorhizal plants. Can. J . Bot. 61, 2924-2929. Torrey, J. G. and Racette, S. (1989). Specificity among the Casuarinaceae in root nodulation by Frankia. Plant Soil 118, 157-164. Tranquillini, W. (1979). Physiological Ecology of the Alpine Timberline. Springer-Verlag, New York. Trappe, J. M. (1962). Fungus associates of ectotrophic mycorrhizae. Bot. Rev. 28, 538-606. Trappe, J . M. (1977). Selection of fungi for ectomycorrhizal inoculation in nurseries. Ann. Rev. Phytopath. 15, 203-222. Trinick, M. J. (1969). Identification of legume nodule bacteroids by the fluorescent antibody reaction. J . Appl. Bact. 32, 181-186. Trinick, M. J . (1982). Biology. In: Nitrogen Fixation Vol. 2: Rhizobium. (Ed. by W. J. Broughton), pp. 76-146. Clarendon Press, Oxford. Tsien, H. C., Cain, P. S . and Schmidt, E. L. (1977). Viability of Rhizobium

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natural vegetation systems. 111. Nutrient transfer between plants with mycorrhizal interconnections. New Phytol. 90, 277-284. Wullstein, L. H . , Bruening, M. L. and Bollen, W. B. (1979). Nitrogen fixation associated with sand grain root sheaths (rhizosheaths) of certain xeric grasses. Physiol. Plant. 46, 1-4. Young, J. P. W. and Johnston, A. W. B. (1989). The evolution of specificity in the legume-Rhizobium symbiosis. Trends Ecol. Evol. 4, 341-349. Young, N. R . and Mytton, L. R. (1983). The response of white clover to different strains of Rhizobium trifolii in hill land reseeding. Grass For. Sci. 38, 13-19.

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Mycorrhizas in Natural Ecosystems M. BRUNDRETT

. . . . . ... . . . . .. ...... . ... .. . . ... . . . .... . . . . . . . . . . . . . . . . . .. . , . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . A. Mycorrhizal Fungi . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Edaphic or Climatic Factors and Mycorrhizal Fungi . . . . . C. TheHost Plant . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Plants and Mycorrhizal Fungi . . . . . . . . . . . . . . . . . . . . E. Edaphic/Environmental Factors, Plants and Mycorrhizas . . F. The Ecology of Mycorrhizal Plants . . . . . . . . . . . . . . . . IV. Conclusions.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Appendix 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I. Summary

11. Introduction . . 111. Mycorrhizal Ecology

171 172 173 175 176 194 196 202 235 257 262 271 271

I. SUMMARY There is now ample evidence to support the common assertion that most plants in natural ecosystems have mycorrhizal associations. Information about the worldwide distribution of plants with different types of mycorrhizal associations is used to establish correlations with the major climatic factors (water, temperature) which regulate the distribution of plants, as well as more localized edaphic conditions. Ecological implications of mycorrhizal associations in natural ecosystems and the role of soil or environmental factors, mycorrhizal fungus characteristics or host plant properties alone or in combination are considered. Factors which can influence the occurrence and effectiveness of mycorrhizal associations include (i) root properties (ii) edaphic or climatic factors (iii) soil 0

ADVANCES I N ECOLOGICAL RESEARCH VOL. 21 ISBN 0-12-013921-9

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organisms, (iv) soil disturbance and (v) host-fungus compatibility. More complex ecological topics (involving, the environment, plants and mycorrhizal fungi) that are discussed include (i) mycorrhizal phenology , (ii) factors responsible for varying degrees of mycorrhizal dependency in host plants, (iii) the role of mycorrhizal hyphae in soil (iv) nutrient competition involving mycorrhizal and non-mycorrhizal plants and (v) mycorrhizal interactions involving pollution and other stresses, the rhizosphere, soil properties and allelopathy. The population ecology of mycorrhizal fungi and the influence of their associations on plant population ecology are also considered.

11. INTRODUCTION Mycorrhizas are highly evolved, mutualistic associations between soil fungi and plant roots. The partners in this association are members of the fungus kingdom (Zygomycetes, Ascomycetes and Basidiomycetes, but not protoctistan fungi such as Oomycetes) and most vascular plants (Harley and Smith, 1983; Kendrick, 1985). In the mycorrhizal literature the term symbiosis is often used to describe these highly interdependent mutualistic relationships where the host plant receives mineral nutrients while the fungus obtains photosynthetically derived carbon compounds (Harley, 1989; Harley and Smith, 1983). At least seven different types of mycorrhizal associations have been recognized, involving different groups of fungi and host plants and distinct morphology patterns (see Hadley, 1982; Harley, 1989, Table 1; Harley and Smith, 1983; Read, 1983). The most common associations are (i) vesicular-arbuscular mycorrhizas (VAM) in which zygomycetous fungi produce arbuscules, hyphae and vesicles within root cortex cells, (ii) ectomycorrhizas (ECM) where Basidiomycetes and other fungi form a mantle around roots and a Hartig net between ,root cells, (iii) orchid mycorrhizas where fungi produce coils of hyphae within roots (or stems) of orchidaceous plants and (iv) ericoid mycorrhizas involving hyphal coils in outer cells of the narrow “hair roots” of plants in the Ericales. In this review, hyphae of a mycorrhizal fungus originating from one entry point in roots or one propagule in soil are referred to as colonies, and colonization refers to the degree of root occupation by mycorrhizal fungi. It has often been stated that most plants in ecosystems have mycorrhizal associations, but there have been no attempts to catalogue the evidence which supports this assertion since Kelly (1950) last summarized information about the worldwide distribution of mycorrhizal plants. However, there have been recent literature surveys which consider correlations between mycorrhizal strategies and plant taxonomy (New-

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man and Reddell, 1987; Trappe, 1987) and one concise regional compilation of host plants (Harley and Harley, 1987). A survey of information about the mycorrhizal status of plants occurring in each of the world’s major ecosystems and edaphic communities is provided in Appendix 1 at the end of this review. This compilation provides a summary of our knowledge about the distribution of mycorrhizal associations in natural ecosystems and allows correlations between these distribution patterns and climactic and edaphic factors to be established. Various aspects of the ecology of mycorrhizal fungi and their associations with plants in natural ecosystems are considered in this review.

111. MYCORRHIZAL ECOLOGY Mycorrhizal associations are regulated by features of the host plant and mycorrhizal fungus, as well as by soil conditions and environmental factors (Harley and Smith, 1983; Mosse and Hayman, 1980). Mycorrhizal ecology can be viewed as regions of overlap between one or more of these three factors (Fig. 1) and the discussion here reflects this

\

Host plant

Fig. 1. Mycorrhizal associations result from three-way interactions between mycorrhizal fungi, host plants and environment/soil conditions, as is illustrated by three overlapping regions in this figure. The labelled regions refer to mycorrhizal ecology subject areas that are sections in this review.

The life cycle of a mycorrhizal association [spores, hyphae. old roots. etc I

1. Fungal propagules

survival

disturbance, predation adverse conditions

dispersal dormancy. quiescence * activation - fungi may respond to: environmental conditions time intervals presence of roots or other organisms 3. Roor growth

Lyoung roots required to form arsocrationl regulated by phenology and environmental factors production of soluble or volatile signals,

2. Active soil hyphae mineral nutrient acquisition by fungus microhabitat preferences ? limited saprobic potential ? spread through soil attraction to roots*?, tropic responses? contact with young host mots

-

4. Hyphae on roor surface

-

proliferation on mot surface

* recognition of potential host ?

fungus morphology changes appressoria (VAM). mantle (ECM)

5. Hyphal penetration inro or between roor cells avoidance or tolerance of host defences recognition by host ? (minimal response) further fungus morphology changes 6. Formation of exchange sires * most

pronounced fungus morphology changes highly branched hyphal StNCtllRS arbuscules (VAM),Hartig net (ECW obvious host responses (protoplasm synthesis, etc.)

7. Acrive exchange processes

limited in duration influence of host. fungus or environment. 8. Senescence of hyphal srructures wirhin roors dissorganization of exchange site hyphae fungal resources withdrawn for: storage by hyphae, vesicles within mot (VAM) storage in mantle (ECM) expon into external mycelium [the cycle recommences] 9. Propagule formation resting spores in soil or mot (VAM) sexual spores from mushrooms, truffles (ECM) mycelial strands and sclerotia (ECM) mycelial networks in soil long-lived mots

-

.root conexrootlostdeath due or secondary growth

10. Roor senescence

to:

parasitism or consumption nutrients in mot & fungus structures recycled decomposition food web

-

Fig. 2. This chart illustrates the sequence of events during mycorrhizal formation, senescence and propagule formation.

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structure. Discussion topics also follow the sequence of events presented in Fig. 2, which is a schematic representation of the life-cycle of a mycorrhizal association, as much as possible. The influence of environmental factors on soils or plants alone have been considered in detail elsewhere (Barbour et a l . , 1987; Etherington, 1982; Fitter and Hay, 1987). Most of the discussion which follows concerns VAM and ECM associations because there have been most studied, but other associations types are also considered.

A. Mycorrhizal Fungi Members of the fungus kingdom obtain nutrition from many sources, including decomposition of organic substrates (saprobes), predation and parasitism and involvement in mutualistic associations (Christensen, 1989; Kendrick, 1985). Many soil fungi are saprobes with the enzymatic ability to digest organic substrates of varying degrees of complexity, but some subsist on very low levels of organic or inorganic substrates (Wainwright, 1988). Mycorrhizal fungi are a major component of the soil microflora in many ecosystems (Harley, 1971) and usually have limited saprobic abilities (Section 111. F. 1). Endophyte properties which would help determine the effectiveness of mycorrhizal associations include the amount of soil hyphae produced relative to root colonization, the rate of hyphal growth and root colony initiation, and physiological characteristics which regulate nutrient absorption or nutrient translocation by hyphae and exchange with the host (Kottke and Oberwinkler, 1986; Smith and Gianinazzi-Pearson, 1988). Fungi forming VAM associations include about 150 species belonging to the genera Gigasporu and Scutellispora (Gigasporaceae), Glomus and Sclerocystis (Glomaceae) and Acnulosporu and Enrrophospora (Acaulosporaceae) in the Zygomycete order Glomales (Morton and Benny, 1990). Mycorrhizal fungi apparently can occupy a particular habitat for thousands of years with little genetic change (Trappe and Molina, 1986) and fossil evidence suggests that VAM associations have been present throughout most of the history of vascular plants (Pirozynski and Dalpe, 1989; Stubblefield and Taylor, 1988). The relatively small number of extant species and the scarcity of sexual reproduction in this group of fungi also suggest that the potential for genetic change within these species is limited (Morton, 1988; Tommerup, 1988). Morton (1990) has proposed that most of the evolution of species in the Glomales occurred early in their evolutionary history and that they have changed relatively little since then. However, the hyphae and spores of VAM fungi are multinucleate and probably also heterokaryotic, so

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genetic changes may occur through hyphal anastomois or somatic recombination involving different nuclei (Tommerup, 1988; Trappe and Molina, 1986). Careful taxonomic studies (Morton, 1988), the use of new methods such as isoenzyme analysis (Hepper et al., 1988) and variable responses to soil and environmental conditions (Section 111. B.5) have demonstrated considerable variation within currently defined taxa of VAM fungi. A wide range of fungi, including thousands of species, belonging to 25 families of Basidiomycetes, seven families of Ascomycetes and a single Zygomycete family, can form ECM associations (Miller, 1982a). Some hosts form associations with many fungi while others are more specific (Duddridge, 1987; Trappe, 1962). Host preferences, metabolic diversity and responses to habitat conditions provide evidence that considerable diversity occurs within this group of fungi (Section 111. B.6). Debaud et al. (1988) found that homokaryotic mycelia of the fungus Hebeloma cylindrosporum produced ECM associations that were similar to those produced by hyphae of the dikaryotic parent. Unfortunately, little is known about the genetics (cytology, nuclear behaviour, mating systems, etc.) of ECM fungi (Harley and Smith, 1983; Trappe and Molina, 1986). The ericoid mycorrhizal fungi that have been identified include Ascomycetes such as Hymenoscyphus ( Pezizella), Myxotrichum and Cymnascellu (Dalpk, 1989; Read, 1983). Fungi forming mycorrhizal associations with orchids include many Rhizoctonia anamorphs (some of which have known teliomorphs), as well as other fungi that may form specific or non-specific associations with their hosts (Currah et al., 1987; Ramsay et al., 1987; Warcup, 1981, 1985).

B. Edaphic or Climatic Factors and Mycorrhizal Fungi 1. Propagules of Mycorrhizal Fungi The spread to new roots, long-range dispersal and persistence of mycorrhizal fungi in ecosystems is dependent on the formation of propagules and their interactions with soil and environmental conditions (Fig. 2). It is important that infective propagules are present when root growth activity occurs, since roots have a limited period of susceptibility (Brundrett and Kendrick, 1990a; Hepper, 1985) and rapid colonization of the root system is required for an effective association (Abbott and Robson, 1984a; Bowen, 1987). The activity of VAM fungi in soils is usually quantified by measuring mycorrhizal formation within roots with a microscope using a clearing and staining procedure (see Brundrett ef a[., 1990; Kormanik and McGraw, 1982). When these methods are used

177 to quantify propagules of mycorrhizal fungi it is important to consider the total length as well as the proportion of host root occupied by VAM colonies (Abbott and Robson, 1991a). The number of successful root contacts (ECM short roots, or individual VAM colonies within roots (which are often called infection units) is the best criterion to measure the infectivity of soil fungi (Huisman, 1982), but hyphal spread along roots and colony extension within roots also contributes to total colonization levels (Sanders and Sheikh, 1983; Smith and Walker, 1981). Increased soil hyphae activity often follows association establishment so it is also important to grow bait plants in the soil being tested for only a short period (several weeks) before enumeration. Propagules of VAM fungi include spores, root fragments containing hyphae and vesicles (storage structures) and soil hyphae (Biermann and Lindermann, 1983; Manjunath and Bagyaraj, 1981; Tommerup and Abbott, 1981). The large soil-borne spores of VAM fungi, are considered by many to be the most important type of inoculum, but their numbers are often poorly correlated with mycorrhizal formation in soils (Abbott and Robson, 1984a, 1991a; Ebbers et af., 1987; McGee, 1989; Mukerji and Kapoor, 1986; Schmidt and Reeves, 1984). Spore production by VAM fungi is influenced by many factors including the host plant (Section 111. D . l ) , and soils in ecosystems often contain low numbers of living spores (Brundrett and Kendrick, 1988; Gay et al., 1982; Janos, 1980b; Read et af., 1976; Schenck and Kinloch, 1980). Living spores of VAM fungi present in the soil may not function as propagules, if they are quiescent (inactive because soil conditions are unsuitable) or have an innate period of dormancy -mechanisms which may help them survive periods of adverse soil conditions (Tommerup, 1987). Melanins (fungal pigments) within their relatively thick walls may help protect the contents of VAM spores, which sometimes receive further protection by forming within structures such as old seed coats (Daniels Hetrick, 1984). Spores are generally considered to be more resistant to adverse conditions than other VAM fungus propagules (Abbott and Robson, 1990). A poor correlation between spore numbers and mycorrhizal initiation, and the rapid initiation (within a few days) of VAM infection that often occurs in ecosystems suggest that a pre-existing network of soil hyphae is often the main source of VAM inoculum (Birch, 1986; Jasper et al., 1989a; McGee, 1989; Powell, 1977; Read et af., 1976). Fragments of dead roots present in the soil can also initiate VAM, provided they are in close proximity to the new roots (McGee, 1987; Rives et a[., 1980), which would happen if they occupy the same soil channels (Went and Stark, 1968). The vesicles produced by many VAM fungi (hyphal swellings containing cytoplasm and lipids) are considered to function as MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

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temporary storage organs, but often elaborate multilayered walls like spores. Vesicles can function as propagules when isolated from roots (Biermann and Lindermann, 1983). Some plants in a deciduous forest community were found to have roots which maintained a living cortex for 2-10 years without undergoing secondary growth and still contained inactive hyphae and vesicles of VAM fungi (Brundrett and Kendrick, 1988). These species with long-lived roots may function as keystone mutualists (Gilbert, 1980), benefitting all host plants by allowing VAM fungi to perennate within them (Brundrett and Kendrick, 1990a). Coarse soil organic matter colonized by VAM fungus hyphae can also contribute to their survival (Warner, 1984) and function as propagules (Koske, 1987b; Nicolson, 1960). There is evidence that some endophytes do not produce spores (Johnson, 1977; McGee, 1989; Powell, 1977), or fail to survive in dried root fragments (Tommerup and Abbott, 1981). The relative importance of spores, old roots and soil hyphae as propagules apparently varies between species of VAM fungi occurring in the same site (McGee, 1989). Since the most important propagules of VAM fungi in soils are generally unknown, it is best to find some measure of the total infectivity of these fungi. The mycorrhizal infectivity of soils can be estimated by most probable number methods (serial dilutions using increasing proportions of sterilized soil), or “bioassays” (where the degree of colonization of a bait plants are measured), but it can also be difficult to interpret results obtained by these methods (see Abbott and Robson, 1991a). Mycorrhizal fungus activity, measured by the presence of mycorrhizal roots and spores, is generally thought to be concentrated near the soil surface, but propagules can be more numerous at greater depths (up to 2-4 m) in arid ecosystems (Virginia et al., 1986; Zajicek et af., 1986). Propagules of ECM fungi include hyphae. mycelial strands and rhizomorphs (Ogawa, 1985; Read et al., 1985), basidiospores (Bowen and Theodorou, 1973; Fox, 1983, 1986b), sclerotia (Fox, 1986a; Gibson et al., 1988) and probably also mycorrhizal roots, but these fungi typically do not produce asexual (conidial) spores (Hutchinson, 1989). Boreal forest soil and leaf litter contains basidiospores which can initiate mycorrhizas (Amaranthus and Perry, 1987; Parke et af., 1983, 1984). Localized patterns of ECM fungus proliferation depend on the production of hyphae, mycelial strands, or rhizomorphs by a particular endophyte (Ogawa, 1985). Hyphal strands of some ECM fungi will only initiate new mycorrhizas if attached to living host roots (Fleming, 1984). Roots with ECM usually live for one or more years and are protected by mantle hyphae (Harley and Smith, 1983), suggesting that they may be important perennating structures.

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2. Dispersal of Mycorrhizal Fungi The spread of mycorrhizal fungi can occur by active processes (hyphal growth through soil) or passive dispersal mechanisms (Daniels Hetrick, 1984). Hyphae of VAM fungi radiate outward from mycorrhizal plants and thus can slowly spread the association to adjacent plants (Warner and Mosse, 1982; Scheltema ef al., 1985b). Dispersal mechanisms are responsible for introduction of mycorrhizal fungi to new geographic locations and the transfer of genetic information. The large (for a fungus) spores of VAM fungi can be suspended in moving air currents (Tommerup, 1982) and wind dispersal has been observed in arid ecosystems (Allen, 1988; Warner ef af., 1987). Transportation of spores by water erosion and human activities (transport of soil and living plants) probably also occurs (Daniels Hetrick, 1984; Walker, 1988). Koske and Gemma (1990) observed that VAM fungus spores produced in rhizome leaf sheaths or quiescent fungal structures within old roots could function as inoculum, even after exposure to sea water. They suggest that this provides a mechanism for the dispersal of mycorrhizal fungi along with fragments of plants which occupy early successional coastal habitats. The ingestion and subsequent defaecation of spores by animals can result in the introduction of VAM into new locations, as was observed during the revegetation of Mt St Helens (Allen, 1988). Trappe and Maser (1976) observed that VAM spores remaind viable after passage through the digestive tract of a rodent. Animals which probably transport VAM fungus spores include small mammals, grasshoppers, worms, ants, wasps and birds (see Table 1). Macroarthropod detritivores such as woodlice (Isopoda) and millipedes (Diplopoda) ingest and disperse mycorrhizal inoculum and may in turn be eaten by predatory beetles (Rabatin and Stinner, 1988). Earthworms, which frequently ingest VAM fungi and expel their spores in casts, are eaten by many small animals which may thus act as vectors for the mycorrhizal fungi (Rabatin and Stinner, 1988). Mycorrhizal fungus spores have been found within organisms which may act as vectors that belong to many trophic levels, but the distances involved and the importance of these dispersal mechanisms in ecosystems is usually not known. Many fungi forming ECM associations have large fruiting structures (mushrooms) that produce abundant quantities of wind-borne spores, but survival and dispersal of these spores may be limited (Bowen and Theodorou, 1973; Harley and Smith, 1983). Sclerotia, including those produced by ECM fungi, can be moved in spring run-off water by floating or adhering to organic material (Malloch ef af., 1987). Some ECM fungi produce hypogeous fruiting bodies that are excavated and

Table 1 Organisms associated with mycorrhizal fungi

Organism Bacteria bacteria-like organelles spiroplasma-like organisms Actinomycetes Actinomycetes Protoctistan fungi (Chytrids) Rhizidiomcopsis sp. Spizellomyces sp. Protoctistan animals (Amoebae) Saccamoeba and Gephramoeba sp. True fungi Anguillospora and Humicola sp. various fungi various fungi Stachybotrys chartarum unidentified fungi *‘Rhizoctonias” aphyllophoralean fungi various “fungicolous” fungi Nematodes Aphelenchoides sp. fungus-feeding nematodes Meloidodera sp. Aphelenchus sp.

Mycorrhizal fungus

Type of association‘

VAM VAM VAM ECM

References

Macdonald and Chandler (1981) Tzean et al. (1983) Ames et al. (1989) Richter et al. (1989)

VAM VAM

3 3

Ross and Ruttencutter (1977) Daniels and Menge (1980)

ECM

4

Chakraborty et al. (1985)

VAM VAM ECM VAM VAM VAM ECM ECM

3 2 2 2,s

Daniels and Menge (1980) Secilia and Bagyaraj (1988) Summerbell (1989) Siqueira et al. (1984) Ross and Ruttencutter (1977) Williams (1985) Fries and Swedjemark (1985) Barnet, 1964; Hawksworth (1981)

Ericoid VAM ECM ECM

3 2,3 3 2.6

Shafer et al. (1981) Ingham (1988) (review) Zak (1967) Sutherland and Fortin (1968)

Arthropods mites collembola (springtails) Pemphigus sp. (aphid) Tetruneuru sp. (aphid) Coleoptera (25+ sp.) Onicus sp. (sowbugs) Grylfus sp. (crickets) ants and wasps Earthworms Lumbricus sp. Birds Hirundo sp. (barn swallow) Animals squirrels, other small mammals and marsuipals

all? VAM

4 4,7

ECM VAM ECM VAM VAM VAM

8 8 6,9 7,lO 7,lO 10

see text Moore et uf. (1985), Finlay (1985), Rabatin and Stinner (1988) Zak (1965) Reddy and Sharma (1981) Fogel and Peck (1976) (review) Rabatin and Stinner (1988) Rabatin and Stinner (1988) McIlveen and Cole (1976)

VAM

7, 10

McIlveen and Cole (1976)

VAM

10

McIlveen and Cole (1976)

ECM VAM

7,9, 10

Blaschke and Baumler (1989) Maser and Maser (1988), McIntire (1984), Malajczuk et ul. (1987), Cowan (1989)

1: found in cytoplasm (role unknown): 2: associated with fungus: 3: occupy or attack spores (parasitic or necrotrophic); 4: feed on hyphae; 5 : inhibit spore germination; 6: feed on eigeous sporphores; 7: ingest spores; 8: feed on mycorrhizal roots; 9: feed on hypogeous sporophores; 10: move spores with soil.

182

M . RRUNDRETT

consumed by small mammals or marsupials and thus spread to new locations (Table 1). Spores of ECM fungi contained in animal faeces can be a viable source of inoculum (Kotter and Farentinos, 1984; Lamont el al., 1985). In western North America, hypogeous fungi form a major part of the diet of squirrels, which in turn are beneficial to the community by transporting nitrogen-fixing bacteria and fungal spores which can establish new mycorrhizal fungus colonies or transfer genetic material to existing colonies (Maser and Maser, 1988). Similar tree-mycorrhizal fungus-dispersing animal inter-relationships also occur in Europe, Australia and New Zealand (Blaschke and Baumler, 1989; Malajczuk et al., 1987; Cowan, 1989). The supply of mycorrhizal inoculum could be limited in some recently created or disturbed habitats if these fungi were less readily dispersed than their host plants (Section 111. F.4).

3. Soil Organisms Associated with Mycorrhizal Fungi The wide variety of mycophilous (Barnet 1964, Hawksworth 1981) soil organisms that are known to ingest, inhabit, or associate with hyphae or spores/sporophores of mycorrhizal fungi are listed in Table 3. This table contains members of most of the trophic levels of organisms which feed in or on soils. Rhizosphere micro-organisms associated with mycorrhizal roots are considered in Section 111. E.3. Some organisms listed in Table 1 may be associated with moribund hyphae or spores and thus have little effect on mycorrhizal associations, but others apparently kill living hyphae and spores of these fungi. Chytrids may occupy VAM fungus spores that already are moribund (Paulitz and Menge, 1984), but in other cases contamination by these organisms is correlated with reduced spore germination (Sylvia and Schenck, 1983). Fungivorous mites are an important component of soil food webs (Coleman, 1985) and are also likely to be major consumers of mycorrhizal fungi (St John and Coleman, 1983). Dense mats of ECM roots and fungi in forest soils can have substantially higher populations of microbes and microarthropods than other areas (Cromack et al., 1988). Fungi that may be parasitic on ECM roots have been observed to form hyphae in the mantle and haustoria within root cells (Agerer. 1990; Brundrett et a[., 1990). Several mycorrhizal root associates, including root inhabiting fungi, nematodes and an aphid-fungus partnership were observed during a study of Canadian forest tree roots (Brundrett et al., 1990) and it is certain that many more interesting mycorrhizal root-soil organism associations will be unearthed. Because of the widespread occurrence and abundance of mycorrhizal fungi in soils, we would expect them to be an important food source for many soil organisms (Daniels Hetrick, 1984; Harley, 1971; Odum and

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Biever, 1984). Fungus-feeding nematodes and springtails have been observed to reduce the efficacy of mycorrhizal associations in some cases (Warnock er al., 1982; Finlay, 1985; Ingham. 1988; Rabatin and Stinner, 1988). Differences in the susceptibility of mycorrhizal fungi to antagonistic microbes has been reported to influence the outcome of competition between endophytes (Godfrey, 1957; Ross and Ruttencutter, 1977). Spores of VAM fungi isolated from soils in natural ecosystems frequently show signs of predation (Brundrett and Kendrick, 1988; Berliner and Torrey, 1989; Janos, 1980a; Koske, 1988) that may be partially responsible for seasonal fluctuations in their abundance (Section 111. E.l). Hyphal grazing by soil organisms has the potential to significantly reduce the efficacy of mycorrhizal associations in ecosystems- by limiting the transport of mineral nutrients to roots (Finlay, 1985; Ingham, 1988; McGonigle and Fitter, 1988b; Rabatin and Stinner, 1988), but also allows nutrients contained in hyphae to be recycled (Coleman, 1985; Perry ef al., 1987).

4. Soil Disturbance and Mycorrhizal Fungi Mycorrhizal propagules can be severely influenced by damage to vegetation and soils resulting from natural processes or human intervention. These destructive processes include intense fires (Dhillion et al., 1988; Klopatek et al., 1988; Wicklow-Howard, 1989), exposure of subsurface soil by erosion (Day et al., 1987; Habte, 1989), or by burrowing animals (Koide and Mooney, 1987) and volcanic activity (Allen, 1988). Human activities which influence mycorrhizas include; topsoil disturbance and stockpiling during mining (Danielson, 1985; Jasper et al., 1987; Rives et al., 1980; Stahl et al., 1988) and clear-cut logging (Janos, 1987; Parke et al., 1984; Perry et al., 1987). Agricultural practices such as tillage (Evans and Miller, 1988), long fallow periods (Thompson, 1987), soil compaction (Wallace, 1987), growth of nonmycorrhizal crops (Section 111. F.3.a), biocide application (Haas et al., 1987; Medve, 1984), or flooding for rice culture (Ilag et al., 1987; Nopamornbodi et al., 1987) can also adversely influence mycorrhizas. These processes usually result in greatly reduced VAM or ECM fungi formation. Forest decline associated with air and/or precipitation borne pollution has become a serious problem in Europe and North America (Klein and Perkins, 1988; Smith, 1990) and various forms of pollution can inhibit mycorrhiza formation in experimental systems (Section 111. E.6). Small scale disruptions due to frost action, shrinkage of drying soil, or the activities of soil animals may also be detrimental to mycorrhizal hyphae in soil (Read and Birch, 1988). Klopatek et a l . (1988) found that high soil temperatures resulting from a fire were correlated with reductions in subsequent VAM formation.

184

M . UKUNDRETT

Soil disturbance may have a direct impact on propagules of mycorrhizal fungi or an indirect affect through changes to soil properties. Disturbance impacts on VAM fungi that have been hypothesized include; (i) a reduction numbers of viable spores, (ii) loss of a hyphal network in the soil, or (iii) the prevention of hyphal growth from root inoculum to new roots (Evans and Miller, 1988; Jasper et al., 1989abc; Rives et al., 1980). The relative importance of these mechanisms has not been fully established and may vary in different situations. A reduction in VAM endophyte diversity from 11 to one species occurred after disturbance in an alpine area (Allen, et al., 1987). Spores of these fungi are fairly resistant to physical and chemical stress (Tommerup and Kidby, 1980) but apparently are killed or removed by severe disturbance. During fallow periods or when soil is stockpiled, the infectivity of spores could decline as their population structure becomes older (without host plants to support new spore production) and as a result of premature germination (Section 111. E . l ) . Daft et LII. (1987) found that spores were still viable after topsoil storage for 12 weeks if temperature and humidity were moderate. They found that spores were more resistant to these conditions than root fragments and much more resistant than hyphae. A network of mycorrhizal fungus hyphae in soil is considered to be an important source of inoculum in natural ecosystems (Section 111. B . l ) . Disruption of this network can result in a substantial loss of mycorrhizal infectivity (Evans and Miller, 1990; Jasper et al., 1989a. b), or reduce the vigour of mycorrhizas that do form-perhaps because they were no longer connected to a common resource pool (Read & Birch 1988). Soil structure changes resulting from disturbance could disrupt the spatial association between old and new roots, or otherwise reduce the effectiveness of root inoculum (Evans and Miller, 1988; Rives et ul., 1980). Changes to soil properties and populations of soil organisms which occur during soil stockpiling can reduce the efficacy of surviving VAM endophytes (Abdul-Kareem and McRae, 1984; Stahl et al., 1988; Waaland and Allen, 1987). Loss of ECM inoculum can occur rapidly in the absence of host plants after clear-cutting, especially with burning and scarification and soil property changes can prevent the reestablishment of surviving fungi (Amaranthus and Perry, 1987; McAfee and Fortin, 1986, 1989; Parke et id., 1984: Perry et al., 1987). During revegetation "early stage" mycorrhizal fungi (Section 111. B.6) such as Pisolithiis tinctorius, which are well adapted to growth in mineral soils, become dominant until forest soil conditions develop (Danielson, 1985; Gardner and Malajczuk, 1988; Stahl Pt ctl., 1988). During the recovery process a succession involving increasing numbers of ECM fungi occurs as soil organic matter builds up (Section 111. B.6). After disturbance, surviving ECM fungus inoculum

M Y C O R R H I Z A S IN NATURAL ECOSYSTEMS

185

may be concentrated in localized soil pockets high in organic materials (Christy et al., 1982: McAfee and Fortin, 1989). Following ecosystem disturbance, a reduction in mycorrhizal inoculum apparently is often responsible for reduced mycorrhiza formation. The level of inoculum in disturbed soil will depend on initial propagule concentrations, the resilience of propagules. their reintroduction by dispersal mechanisms and on the impact of changes to soil conditions on the activity of surviving fungi. The mycorrhizal dependency of recolonizing vegetation also varies (Section 111. F.4). In some cases it may be advantageous or necessary to introduce mycorrhizal fungi to replace lost inoculum or, that provide a greater benefit than indigenous fungi (Danielson, 1985; Cook and Lefor, 1990; Jasper et al., 1988; Perry et al., 1987; Stahl et al., 1988).

5. Climatic or Edaphic Specificity of Mycorrliizal Fungi Environmental factors and soil conditions influence the occurrence of mycorrhizal associations in ecosystems (Section 111. F.S), but it is hard to examine the direct impacts of these factors on mycorrhizal fungi because they rarely occur in nature without a host and members of the Glomales can not be grown axenically (Harley and Smith, 1983). However, some evidence of the physiological diversity of mycorrhizal fungi has been provided by comparing experimental responses to soil pH, soil nutrient levels, soil moisture, salinity, temperature and other factors (Abbott and Robson, 1991a; Daniels Hetrick, 1984; Morton, 1988; Slankis, 1974; Trappe and Molina, 1986). Most of this data has been collected using simplified experimental systems which allow the influence of one factor on one mycorrhizal fungus to be examined, but some field data are also available for comparison. Changes to soil properties occurring during succession or between sites with similar climates can be correlated with the predominance of different species or isolates of VAM fungi (Bethenfalvay et al., 1982; Gerschefske Kitt et al., 1987; Puppi and Reiss, 1987; Rose, 1988). There is limited evidence that climatic factors can influence the distribution of mycorrhizal fungus taxa (Section 111. B.6). Ebbers et al. (1987) and Anderson et al. (1984) discovered changes in predominate species of VAM fungi across a soil moisture (soil fertility) gradient in a prairie site, which had a much greater influence on plant populations. Henkel et al. (1989) observed that isolates of four VAM fungi from adjacent ridgetop, mid-slope and basal sites in an arid plant community were most infective in the soil from which they were collected and less infective in soil from the other two sites. They suggested that these isolates had adapted to phosphorus levels or other factors in the soil where they occur. Adelman and Morton (1986), Graham et al. (1982b),

186

M. B R U N D R E I T

Molina et al. (1978), Porter er al. (1987b) and Stahl et al. (1988) also observed that clonal isolates of VAM fungi were more effective when used in their native soil type. Further evidence of the physiological diversity of VAM fungi is provided by comparing responses of different species or isolates to physical conditions (Table 2). These comparisons have demonstrated variations between taxa and intraspecific variability within species of VAM fungi in their ability to promote plant growth when exposed to the factors listed in Table 2. These fungi apparently have a limited tolerance range to environmental conditions (Stahl er al., 1988) and possess specific adaptations to the soil in which they occur (Lambert er al., 1980). These adaptations apparently can influence the outcome of competition between VAM fungi (Gerschefske Kitt et al., 1987). The effect of low soil pH on VAM associations is discussed by Howeler et al. (1987) and Robson and Abbott (1989). Some endophytes can still provide substantial benefits to the host plant in soils with low pH and high aluminium levels, while others are less effective (Table 2). It has sometimes been observed that the fine endophyte called Glomus renue is more abundant in acidic soils (Gianinazzi-Pearson et al., 1980; Wang et al., 1985). Porter er al. (1987a) found the distribution of VAM fungus taxa in Western Australia to be highly correlated with soil pH. Bethlenfalvay er al. (1989) proposed that the term "edaphotype" be used to describe intraspecific variants of mycorrhizal fungi isolated from different soils that differ in their physiological response to various conditions. There are many statements in the mycorrhizal literature about the physiological, ecological, or mutualistic characteristics of species of VAM fungi that actually only describe one particular clonal isolate (Morton, 1990). Unfortunately, VAM research has been concerned with plant responses to mycorrhizas with little consideration of specific endophytes, thus creating the impression that these fungi are functionally equivalent (Abbott and Robson, 1991a; Morton, 1988; Walker, 1988). This impression has been strengthened by the classification problems of fungi such as Glonius fasciculatum which originally had a worldwide distribution and was accredited with remarkable genetic and physiological plasticity (Morton, 1988); this fungus has now been more precisely defined (Walker and Koske, 1987). It is probable that many of the earlier reports of the occurrence of Glomus fasciculatum and experiments conducted with this fungus refer to other, as yet unnamed fungi. If the taxonomy problems associated with many of the VAM fungi originally considered to have wide geographical ranges and the physiological variations that often occur within species of these fungi are considered, it becomes apparent that these fungi do exhibit considerable

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

187

Table 2 Evidence of the adaptation of mycorrhizal fungi to localized soil conditions as provided by comparisons between species or geographic isolates of species in responses to various conditions Soil or environmental condition VAM fungi high or low soil P levels

soil micronutrient levels aridity salinity low soil pH

toxic levels of metals low or high temperatures ECM fungi water stress calcium availability different pH levels toxic levels of metals high or low temperatures

References Boerner (1990), Davis et al. (1984), Henkel el al. (1989), Haas and Krikun, 1985), Johnson (1976), Krikun (1983), Thomson et al. (1986) Killham (1985) Bethlenfalvay et al. (1989), Sieverding and Tor0 (1988), Simpson and Daft (1990a), Stahl and Smith (1984) Pond et al. (1984) Adelman and Morton (1986), Hayman and Tavares (198S), Howeler et al. (1987), Koomen et al. (1987), Porter et al. (1987b) Dueck et al. (1986), Gildon and Tinker (1983) Koslowsky and Boerner (1989) Dodd and Jeffries (1986), Raju et al. (1990) Schenck and Smith (1982), Sieverding (1988) Coleman et al. (1989), Parke et al. (1983), Slankis (1974) Lapeyrie and Bruchet (1986) Hung and Trappe (1983), McAfee and Fortin (1987) Denny and Wilkins (1987), Jones and Hutchinson (1988), Thompson and Medve (1984) Cline et al. (1987), Gibson et a l . (1988), Ingleby et al. (1989, Slankis (1974)

physiological diversity and probably have evolved specific adaptations to the edaphic and environmental conditions of the sites where they occur. The distribution and mycorrhizal efficacy of fungi forming ECM associations is also influenced by climatic and edaphic factors (Bowen and Theodorou, 1973; Slankis, 1974; Harley and Smith, 1983). These fungi are generally considered to be acidophilic (preferring a low soil pH) inhabitants of litter layers near the soil surface (A horizon), but some “early stage fungi” prefer calcareous mineral soils (Section 111. B.6). Isolates of ECM fungi show considerable inter- and intraspecific variations in responses to the factors listed in Table 2. Most of these results were obtained from in uirro experiments which can be poorly

188

M . BRUNDRETT

correlated with responses to similar factors in soils (Cline et al., 1987; Coleman et a[., 1989; Hung and Trappe, 1983). It has been suggested that variations in tolerances to physical factors may be responsible for soil specificity (Last et al., 1984) and restricted geographic ranges (Ingleby et al., 1985), and may influence the outcome of fungal competition (McAfee and Fortin, 1987) or host responses to factors such as drought (Parke et al., 1983). It is apparent that a wide range of variations in tolerance to edaphic and climatic factors (such as temperature extremes, drought, soil toxicity etc.) often occur both between and within species of mycorrhizal fungi and these variations may represent adaptation to specific site conditions by poorly understood genetic mechanisms (Morton, 1990; Tommerup, 1988; Trappe, 1977; Trappe and Molina, 1986).

6 . The Population Ecology of Mycorrhizal Fungi Detailed surveys of VAM fungus spores have provided information on the numbers of endophyte species occurring in some plant communities (Table 3). From these surveys it can be seen that the soil in one location normally contains more than one VAM fungus and may contain a fairly wide diversity of these fungi (considering that only about 150 species are known-Morton 1990). It is not possible to use the results in Table 2 to examine the influence of soil conditions or environmental factors on VAM fungus diversity, because of differences in sampling methodologies (sampling intensity and the area surveyed vary considerably) and the limited number of surveys involved). Several VAM fungi can occur within a single root (Abbott and Robson, 1984b, 1991a). It is possible to identify particular endophytes by comparing their colonization patterns within roots (Abbott and Robson, 1982), or using recently devised immunological or biochemical techniques (Hepper et al., 1988; Morton, 1988; Rosendahl et al., 1989). However, in most cases investigators measure overall root colonization levels without considering the relative contribution of particular endophytes, so our knowledge of the ecology of VAM fungi is based on the occurrence of their spores. Spore occurrence data may be misleading because spore abundance can be poorly correlated with mycorrhiza formation (Section 111. B. l ) , fungi which do not produce recognizable spores may be present (Johnson, 1977; McGee, 1989; Morton, 1988) and roots of different plants often intermingle so spores could occur under plants that they were not associated with. The spore characteristics used to identify species of VAM fungi are thought to be relatively conservative features so that considerable genetic variations may well occur within a species of VAM fungus (Morton, 1990) and this intraspecific diversity may also have to be considered.

Table 3 Populations of VAM fungi occurring at various sites

VAM species

Habitats sampled ( n )

Location

References

9- 18 11-18 2-12 9 18

Coastal sand dunes (1) Coastal sand dunes (6) Coastal sand dunes (4) Coastal sand dunes (1) Arid mountains (region) Arid grassland (3) Grasslands (27) Prairie (1) Wheat fields (4) Mediterranean shrubland (1) Pasture, crop, or native (20) Native cerrado vegetation (1) Field crops (4) Forest (1) Field crops (6) Forest (1) Pasture or forest (37) Seminatural or agricultural (4) Agro-forestry (2) Apple orchard (1)

Eastern USA Eastern USA Hawaii, USA Florida, USA Arizona. USA Wyoming, USA Western USA Kansas, USA

Tews and Koske (1986) Koske (1987a) Koske (1988) Sylvia and Will (1988) Bloss and Walker (1987) Henkel et al. (1989) Molina et a l . (1978) Daniels Hetrick and Bloom (1983)

Southern Australia Western Australia Brazil

McGee (1989) Abbott and Robson (1982) Schenck et al. (1989)

Florida, USA

Schenck and Kinloch (1980)

New Zealand New Zealand Southern USA Iowa, USA Quebec. Canada

Johnson (1977) Powell (1977) Adelman and Morton (1986) Walker et af. (1982) Dalpe et al. (1986)

4

2-5 21 3-6 6 3-5 18 7-15 10 7- 12 10 1-5 3-8 10-12 7

190

M . BRUNDRETT

The occurrence of several VAM fungi in soils or within roots suggests that interspecific competition between them is possible. Koske (1981) could find no evidence that any of the VAM endophytes occurring in a coastal ecosystem were better competitors than others and considered environmental factors and host plants to be more important factors influencing their distribution. However, Gemma et al. (1989) observed seasonal variations in spore production between co-existing endophytes and the fact that abundant spore production by one VAM fungus was usually correlated with lower levels of spore production by others, which may have been due to antagonism between species. When several isolates of VAM fungi are inoculated together in pot culture experiments, some have proven to be better competitors than others (Lopez-Aguillon and Mosse, 1987; Van Nuffelen and Schenck, 1984; Wilson, 1984). In these kinds of studies the most successful fungi generally were those that colonized roots most rapidly (Abbott and Robson, 1984b; Wilson, 1984). The outcome of competition between endophytes would be expected to depend on the placement and amount of their inoculum, their hyphal growth rates in soil and interactions within roots (Abbott and Robson, 1991a,b; Hepper e f al., 1988). Root elongation rates of the plants in natural ecosystems, for which this has been measured, are substantially slower than those of crop plants (Brundrett and Kendrick, 1990a; Huisman, 1982; Lyr and Hoffmann, 1967; Russell, 1977), so rapid responses to the presence of roots and hyphal elongation rates should be less important in nature. It is possible that mycorrhizal fungi with slow hyphal growth rates would preferentially occupy slowly elongating roots, while faster fungi in the same soil colonize roots with greater rates of elongation. Competition between endophytes occurring in the same soil could be reduced if their activities were chronologically separated by phenological differences, or spatially separated by substrate preferences, or host specificity, but the microhabitat preferences of VAM fungi have rarely been considered. Fungi with narrow hyphae (called fine endophytes or Glomus tenue) can be readily distinguished from other VAM fungi when cleared and stained roots are examined. Dodd and Jeffries (1986) observed that a fine endophyte was most active during the winter, while other endophytes predominated at other times. It might be expected that mycorrhizal fungi which have periods of activity and quiescence would preferentially associate with roots that were active at the same times of the year. McGee (1989) observed that propagules of a fine endophyte were more abundant near the soil surface while other fungi were more abundant at greater depths. Anderson et a/. (1984) found that different VAM fungi were most common at opposite ends of a soil moisture-fertility gradient at a prairie site. Factorial experiments involv-

MYCORRHIZAS IN NATURAL ECOSYSTEMS

191

ing many combinations of host plants and VAM fungi have rarely demonstrated specificity (Section 111. D . l ) , but almost nothing is known about the specificity of interactions between host plants and mycorrhizal fungi in nature. McGonigle and Fitter (1990) have observed that fine endophyte hyphae preferentially colonized roots of the grass Holcus lanatus, while endophytes with coarse hyphae were more common in other plants at the site throughout the year. This report of “ecological specificity” and the other examples discussed above suggest that specialization in the phenology, and microhabitat preferences, and perhaps some degree of host specificity may limit interactions between the isolates of VAM fungi present in a soil. The fungi forming ECM associations include a large group of Basidiomycetes as well as other fungi and it is normal for a wide diversity of these fungi to be present on the roots of one host tree (Mason et al., 1987). While it is possible to identify the ECM fungi associated with roots by anatomical studies (Agerer, 1986) or other means, most of our information about the occurrence of these fungi comes from observations of their above-ground reproductive structures- which are usually closely associated with mycorrhizal roots (Hilton et al., 1989; Trappe, 1962). Villeneuve er al. (1989) observed that the diversity of ECM fungi occurring in several temperate deciduous and coniferous forest sites was positively correlated with the dominance of potential host trees. Bills er al. (1986) observed that ECM fungus diversity was greater in mixed hardwood forests than in coniferous forests where one tree predominated. Tyler (1985) examined the influence of edaphic factors on the distribution of macrofungi in European forest sites (dominated by the ECM tree Fagus). In this study the relative importance of ECM fungi increased (and saprobes decreased) in more acidic soils and the distribution of many fungi was correlated with edaphic factors, such as soil organic matter and metal ion content. Many ECM fungi have a wide host range and will form associations with hosts which originate on the other side of the globe, but others specifically associate with particular host taxa (Duddridge, 1987; Le Tacon et al., 1987; Malajczuk et al., 1982; Molina, 1981; Molina and Trappe, 1982b). Kropp and Trappe (1982) have suggested that plants such as Alnus riihra and Pseudotsiiga rnenziesii. which often form pure stands during early succession, tend to form specific associations with ECM fungi, while species such as Tsinga heterophyila which become established in the shaded understorey of other trees usually have non-specific ECM associates. Fungi that have specific ECM associations will be restricted in their distribution by the occurrence of their hosts, but other fungi may have specialized habitat preferences etc. which influence their occurrence. When ECM fungi are introduced to a site. as

192

M . BRLINDRETT

when inoculated host seedlings are transplanted. their success depends on their ability to spread through the soil to new roots and on the outcome of competition with indigenous fungi (McAfee and Fortin, 1986). When ecosystems recover from severe disturbance a succession of different ECM fungi will associate with host plants growing at a site, beginning with “early stage” and ending with “late stage” fungi (Fleming, 1985; Gardener and Malajczuk, 1988; Mason et al., 1987). ”Early stage” but not “late stage” fungi can be easily introduced in disturbed sites (Danielson, 1985; Fox, 1986b). Physiological differences between these two groups of fungi are apparent in aseptic culture experiments (Gibson and Deacon, 1990). Early stage fungi are associated with young forests where ECM roots occur in mineral soil, while late stage fungi form ECM in the litter layer of mature forest soils (Dighton and Mason, 1985; Mason et al., 1987). Fungi in the former group may be inhibited by changes to soil chemical and physical properties associated with tree leaf litter, etc. which occur during succession (Perry and Choquette, 1987). Gardener and Malajczuk (1988) observed five ECM fungi associated with Eucalyptus seedlings in a l-year-old revegetation site which contrasted with 40 species in a mature Eucalyptus forest. Fungal diversity generally increases until late in succession, when the number of species present may decline because those with more specialized host or substrate preferences predominate (Bills et al., 1986; Mason et al., 1987). Different clones of a mycorrhizal fungus may be present in the same soil and form colonies which expand by directional or radial hyphal growth. When colonies of the same mycorrhizal fungus overlap they may join by hyphal anastomosis (non-reproductive cell fusion). Hyphal anastomosis is a distinguishing characteristic of higher (Basidiomycete and Ascomyete) fungi and is considered not to occur in the Zygomycetes (Kendrick, 1985). but does occur in members of the Glomales (Tommerup, 1988). Anastomosis is an important mechanism by which mycorrhizal fungi can maintain continuity, exchange nuclear and extranuclear genetic information and reproduce sexually (Tommerup, 1988; Trappe and Molina, 1986). We would expect hyphal anastomosis to be common in ECM fungi (most of which are Basidiomycetes or Ascomycetes), but the frequency with which this event occurs between clones of VAM fungi in soil is unknown. Frequent anastomosis would appear to be essential, especially after a period of quiescence when fungal activity may have to be re-initiated form various scattered propagules. Otherwise clones would remain separate and compete for soil resources and host roots with which to form associations (this may occur to some extent). Tommerup (1988) has observed that spores of mycorrhizal fungi

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isolated 5 m apart formed compatible anastomoses (evidence for original continuity), while the compatibility of spores which originated 150 m apart was lower (71%). Fries (1987b) found that many genetically distinct isolates of one species of ECM fungus could occur in a small forest area. The natural distribution of plant species is limited by their tolerance to environmental conditions, especially periods of extremely cold temperatures, or when soils are dry, but biotic and soil factors are also important (Barbour et al., 1987; Woodward, 1987). One might also expect that the distribution of mycorrhizal fungi would also be influenced by similar factors, since it would be impossible for one isolate of these fungi to be well adapted to both high and low extremes in temperature, moisture, pH etc. Some evidence for the climatic adaptation of endophytes is provided by the optimum, maximum and minimum germination temperatures of mycorrhizal fungi (Tommerup, 1983b) which can be higher in isolates from warmer regions than for fungi originally from cooler locations (Schenck et al., 1975). Koske (1987a) found that some VAM fungi were more abundant at the northern or southern end of a 355-km latitudinal (temperature) gradient of sites with similar coastal vegetation. Environmental factors might be expected to override the more localized mycorrhizal fungus distribution influences (soil conditions, preferred host plants, dispersal effects) considered in previous sections. Unfortunately, there has been a tendency to promote the use of one particular fungus isolate (that would have a limited distribution in nature) as if it had universal applicability in mycorrhizal applications. Considerable thought has been given the criteria used to select “superior” isolates of mycorrhizal fungi for field inoculation programmes (Abbott and Robson, 1991b: Howeler et a l . , 1987; Trappe, 1977), but we must also consider the limits to their ecological adaptability if one isolate is to be successfully used in a wide range of habitats/soils. The diversity of ECM fungi present in communities of associated plants is typically much higher than that of VAM fungi (although additional intraspecific genetic diversity may also be important). A diverse assemblage of ECM fungi is present in communities which generally contain relatively low diversity of host plants (Section lII.F.S). It is necessary to envisage a high degree of ecological specialization of ECM fungi (with regards to host, microhabitat, phenology, or substrate preferences), or complex biotic interactions between these fungi to understand why so many of them can co-exist in relatively small volumes of soil. Little is known about the ecology of fungi forming less common types of mycorrhizal associations. Massive destruction of native vegetation is occurring because of

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deforestation in tropical regions (Jordan, 1985) and forest decline in temperate regions (Klein and Perkins, 1988). While mycorrhizal fungi often exhibit little host specificity, evidence is accumulating that they have become adapted to specific edaphic conditions. Substantial losses in fungal diversity occur when ecosystems are disturbed or converted to agriculture (Allen et al., 1987; Daniels Hetrick and Bloom, 1983; Schenck and Kinloch, 1980; Schenck ef al., 1989). These processes are likely to result in a permanent loss of some of the isolates of mycorrhizal fungi which have become highly adapted to local conditions. During subsequent attempts at ecosystem reconstruction, the impact of this reduction in genetic resources of mycorrhizal fungi will depend on how rapidly surviving fungi adapt to changing soil conditions during succession and how effectively well-adapted isolates are dispersed from remnants of native vegetation.

C. The Host Plant

1. The Structural Diversity and Furzctioii of Roots With few exceptions, roots are essential to the growth and survival of plants. As much as 30% of the plant genome is involved in defining root characteristics (Zobel, 1986) and roots are normally responsible for acquiring the resources (water and mineral nutrients) most often limiting plant growth in ecosystems (Chapin el ul., 1986; Fitter, 1986a). The primary functions of roots include (i) absorption of water and mineral nutrients, (ii) anchorage and physical support, (iii) storage and (iv) mycorrhizal formation (Esau, 1965; Russell 1977). Roots may also support symbiotic nitrogen fixing associations (Gibson and Jordan, 1983; Torrey, 1978). influence shoot growth by growth-regulator production (Carmi and Heuer, 1981; Richards and Rowe, 1977) and in specialized cases may be involved in vegetative propagation or parasitic attachment (Esau. 1965). The structural and functional diversity in roots is generally considered to be much lower than that of plant shoots (Fitter, 1987). It is certainly true that roots essentially are elongated cylinders which often appear superficially similar. However, anatomical or chemical variations between the roots of different species can be sufficient to allow their identification in soils collected from natural ecosystems (Brundrett and Kendrick, 1988; Brundrett rt al.. 1990; Chilvers, 1972). Thus with practice it may often be possible to identify roots in mixed samples by examination of their superficial characteristics, or during the course of

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mycorrhizal assessment. Individual roots can pass through three distinct developmental phases: (i) growth, (ii) maturation and (iii) (in some cases) secondary growth. The importance and duration of these stages varies between plants and root system components. Plants produce a number of types (orders) of roots, including tap, lateral, basal and adventitious roots, that are physiologically, structurally and genetically distinct (Zobel, 1986). For example, lateral roots typically are narrower in diameter, grow less rapidly and have shorter lifespans than roots with higher branching orders (Russell, 1977). The fine laterals (feeder roots) of trees, especially those forming ECM, are often heterorhizic -differentiated into long and short elements, while those with VAM usually have more extensive lateral root systems, without heterorhizy (Brundrett et al., 1990; Kubikovfi, 1967). Tree roots belong to four categories resulting from structural differences between Angiosperms and Gymnosperms and those with VAM or ECM associations (Brundrett er al., 1990). In the later group, distantly related trees have evolved similar, heterorhizic roots with epidermal Hartig nets. Correlations between root structure and function in natural ecosystems are hard to establish because of the scarcity of careful morphological investigations and the frequent use of vague expressions such as “brown” or “suberized” roots (which do not distinguish between tannin accumulation, exodermis, periderm suberization, or root deathRichards and Considine, 1981). There is also a paucity of information about root phenology (which could be used to predict when mycorrhizal associations are likely to be active-Section 1II.E.1). Root structural complexity (protective features such as suberized epidermal, hypodermal, or periderm cell walls) is often correlated with long root lifespans (Brundrett and Kendrick, 1988; Pienaar, 1968), or difficult substrates such as alpine soils (Luhan, 1955). Plants which grow in waterlogged soils generally have large air spaces in their cortex which helps them survive in anaerobic soils (Armstrong, 1979; Justin and Armstrong, 1987). Roots occurring in arid regions often have a cortex which collapses to leave an epidermal/exodermal sheath surrounding roots when the soil is dry (Drew, 1987; Hayden, 1919; Ginzburg, 1966). It is probable that many other correlations between root structure and function occur in natural ecosystems and would be discovered if the roots of plants in a variety of habitats were examined. Roots absorb and translocate nutrients by mechanisms that have been well examined in experimental systems (Marschner, 1986), but it may be difficult to apply some of this knowledge to plants in ecosystems because of many complicating factors (Tinker, 1990). These factors include pH changes, root hairs, exudates and bacteria in the rhizosphere (Tinker, 1990).

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Root morphological criteria of relevance to any consideration of root nutrient acquisition and mycorrhizal formation include the presence of an exodermis (a barrier to solutes and microbes (Peterson, 1988)), heterorhizy (long and short roots), the proportion of roots at different developmental stages (primary roots with a cortex, secondary roots with a periderm, or senescent roots) and their duration and differences in root activity (lifespan and growth rate). Interactions between root structure and mycorrhiza formation will be considered in Section III.E.4.b.

D. Plants and Mycorrhizal Fungi The physiology of mycorrhizal associations has been well discussed by Hayman (1983), Harley and Smith (1983) and Smith and GianinazziPearson ( 1988). Mycorrhizal associations are generally considered to benefit host plants by enhancing mineral nutrient acquisition, especially with regards to phosphorus (Section 1II.E.S). Nitrogen supply by ECM and ericoid associations is also considered to be important (Section III.F.1) and VAM associations may also improve nitrogen uptake (Barea et al., 1989). Increase in the absorption of minor nutrients, such as Mg, Cu and Zn have also been observed, but Mn uptake can be reduced (Arines and Vilarino, 1989; Harley and Smith, 1983: Hayman, 1983; Killham, 1985; Pacovsky. 1986). Other less specific changes to host physiology, which include alterations in nutrient requirements, membrane composition and metabolite levels, apparently occur even when nutrient input is negligible (Dehne, 1986; Pacovsky. 1986). Mycorrhizal fungi (ECM and ericoid) apparently can influence host morphology and physiology by producing plant hormones, such as ethylene and auxins, which may be responsible for the reduced apical growth of mycorrhizal short roots (Berta et af., 1988: Gay and Debaud, 1987: Rupp et al., 1989). Root growth is usually only slightly affected by VAM (Section III.E.S.b), but a detrimental reduction in root elongation occurs in some cases (Jones and Hendrix, 1987). Mycorrhizal associations have been implicated in increased host resistance to disease and other stresses (Section III.E.6).

1. Compatibility and Specificity Figure 2 illustrates the series of events which result in mycorrhizal formation. At the beginning of the process, root and mycorrhizal fungus activity is independently initiated and regulated (both partners may be responding to the same soil or environmental conditions), but there is strong evidence of genetic interactions between the mutualistic partners

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in the later stages of this process. Evidence of genome expression changes in the fungal partners is provided by hyphal structure and behaviour at the root surface, but the response by roots apparently is largely restricted to individual cells forming exchange sites (GianinazziPearson, 1984). The widespread susceptibility of plant roots to colonization by mycorrhizal fungi may be explained by specific comparability systems, or because mycorrhizal fungi somehow avoid or fail to elicit host defence mechanisms (Gianinazzi-Pearson and Gianinazzi, 1989). There is little evidence of host-fungus specificity in most types of mycorrhizal associations (Duddridge, 1987; Gianinazzi-Pearson, 1984; Harley and Smith, 1983). Ineffective VAM associations have been discovered in only a few of the many host plant and mycorrhizal fungus combinations tried in synthesis experiments (Johnson, 1977; Giovannetti and Hepper, 1985). Thus relatively few endophytes ( k l 5 0 members of the Glomales) can form associations with most members of the plant kingdom (Morton, 1988, 1990). Genotypic variations within a host species can influence the degree of VAM formation (Azdon and Ocampo, 1981; Krishna et al., 1985; Lackie et al., 1988; Sieverding and Galvez, 1988; Thomas and Ghai, 1987). Some hosts provide more benefit to VAM fungi than others, as is suggested by differences in the magnitude of spore production. but in most cases spore formation is closely related to the total length of mycorrhizal roots produced by a given host (Giovannetti et al., 1988; Daniels Hetrick and Bloom, 1986; Howeller et al., 1987; Pellet and Sieverding, 1986; Simpson and Daft, 1990b; Struble and Skipper, 1988). The adaptation of mycorrhizal fungi to particular soil conditions (Section 111. B.5), apparently is more common than specific interactions with host plants. Thus in experimental systems incompatible host-fungus combinations are rare, but in ecosystems many of these combinations may be less successful because the fungi are poorly adapted to the normal habitat of plants. However, even if environmental and soil conditions could somehow be excluded from consideration, particular endophytes are also likely to exhibit differences in metabolic competence (the ability to obtain and transport nutrients) (Smith and Gianinazzi-Pearson, 1988). McGonigle and Fitter (1990) observed the preferential association between a VAM fungus with fine hyphae and a grass species, but there have been few other attempts to identify the VAM fungus associates of plants in natural ecosystems (Section 111. B.6). The assertion that VAM associations lack host-fungus specificity may well be a reflection of how little we known about these fungi. Observations of the occurrence of above-ground fructifications of ECM fungi has provided much information about associated host plants

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and the geographic ranges of these fungi (Mason et al., 1987). There is usually a high correlation between the occurrence of fruiting structures and mycorrhizal formation by ECM fungi (Gardner and Malajczuk. 1988; Trappe, 1962), but association lists produced in this way are certain to contain some errors. One such erroneous report involved ash trees (Fraxinus sp.), which are known to have VAM associations, and the fungus Boletinellus merulioides, which is now known to associate with root-feeding aphids (Brundrett and Kendrick, 1987). Details of mantle structure have also been used to identify associated ECM fungi in some cases (Agerer, 1986; Mason et al., 1987). Most ECM fungi associate with a broad range of host plants, but incompatible host-fungus combinations have been found and some fungi specifically associate with one genus (or possibly a single species) of host trees (Duddridge, 1987). Clonal variations within Sitka spruce can influence populations of ECM fungi associated with their roots (Walker et al., 1986). The compatibility of host plant-ECM fungus combinations has been tested using artificial conditions (host seedlings grown in aseptic media) and fungi that colonize roots best under these conditions are often those that form sporocarps in close association with the same host in the field (Molina and Trappe, 1982b). Attempts to form associations between ECM fungi and host plants that are incompatible (at least with the experimental conditions used) can result in a “wounding response” (lignification or phenolic accumulation and cell disruption) in the root cortex (Malajczuk et al., 1984; Molina and Trappe, 1982b). There is strong evidence of cellular and genetic interactions between host plants and mycorrhizal fungi (Gianinazzi-Pearson and Gianinazzi, 1989, but these relatively subtle interactions may be hard to separate from environmental/edaphic influences (Section 111. B .6) on the occurrence of mycorrhizal fungi in natural ecosystems.

2. Mycorrhizal Definitions Researchers have used different criteria to designate the mycorrhizal status of plants. The Hartig net-hyphae forming a labyrinthic exchange site between root cells (Kottke and Oberwinkler, 1986; Massicotte et al., 1987) should be used to identify roots with functional ECM associations. In most cases ECM associations are quantified by counting ECM short root tips (using superficial dissecting microscope examinations) and making occasional sections to confirm the presence of a Hartig net (Grand and Harvey, 1982). Unfortunately, some reports of ECM are only substantiated by the presence of fruiting structures produced by putatively mycorrhizal fungi in the vicinity of a plant (Harley and Smith, 1983). There are many early reports of VAM or ECM associations of

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trees which have since been well established to form mycorrhizas of the other type (see Harley and Harley, 1987 for examples). These designations should perhaps be re-evaluated if they were not originally based on careful morphological observations. However, there are well documented cases of trees belonging to taxa which normally have ECM also forming VAM, especially as seedlings (Chilvers et al., 1987; Lodge, 1989; Vozzo and Hacskaylo, 1974) and such dual associations apparently are common in Australian arid communities (Kope and Warcup, 1986; Warcup, 1980, 1988). Samples collected in natural ecosystems will almost inevitably contain roots of more than one species, which must be carefully separated on the basis of their anatomical characteristics to avoid misinterpretations. The presence of arbuscules (the main site of host-fungus nutrient exchange: Harley and Smith, 1983; Smith and Gianinazzi-Pearson, 1988), is normally used to designate VAM associations, but the presence of hyphae and vesicles alone have also been used as evidence. With practice, these latter structures can usually be distinguished from those produced by saprobic or parasitic fungi, but they make unreliable indicators since they also occur in senescent roots of non-host species, rhizome scales, etc. (Brundrett and Kendrick, 1988; Hirrel et al., 1978; Stasz and Sakai, 1984). Arbuscules are ephemeral structures, which may be absent if samples are collected when roots are inactive (Section 111. E . l ) . Thus it may be difficult to designate plants as mycorrhizal if they only have old or senescent roots when harvested. In ecosystem surveys it may be best to define VAM colonization levels as the proportion of a plant's root system that, when susceptible to colonization, supported an active association with arbuscules (Brundrett and Kendrick, 1988). This requires a prior understanding of host root phenology, or the collection of root samples throughout the year. Mycorrhizal surveys independently conducted in diverse localities have shown that members of certain plant families rarely form mycorrhizal associations. These typically non-mycorrhizas families are listed in Table 4, but their have been some reports of mycorrhizas in most of these families (Section 111. E.4). Studies of plants in ecosystems which have found plants within the same genus or family to have consistent mycorrhizal associations greatly outnumber reports of variable mycorrhizal relations within taxa (Brundrett and Kendrick, 1988; Brundrett et af., 1990; Maeda, 1954; Pendleton and Smith, 1983; Selivanov and Eleusenova, 1974; Trappe, 1981, etc.). Host plants forming Ericoid and Orchid mycorrhizas are taxonomically restricted, so these associations usually do not present identification problems (Hadley, 1982; Read, 1983). Members of the Ericales (Erica-

Table 4 Plant families which are predominantly non-mycorrhizal CLASS SUBCLASS Order

Family"

MAGNOLIOPSIDA MAGNOLJIDAE Nymphaeales 5 Families Papaveraccae(?) Papaverales Fumariaceae HAMAMELIDAE Urticaceae( ?) Uri ticales

Habit Accumulated chemicals and habitat" Primitive< Advanced"

Aq Hb Hb

T f P+ E f T- P - E T- P- E-

Hb

T+ P+

Ak+ Ak+ Ak+

CAKYOPHYl.L.IL)AE

Caryophyllales Phytolaccaceae(?) Nyctaginaceae Aizoaceae Chenopodiaceae Amaranthaceae Portulacaceae Caryophyllaceae Polygonaceae Polygonales

D ILLEN IIDA E Lecythidales Nepenthales Capparales Ebenales ROSIDAE Proteales Podostemales Haloragales Rhizophorales Santalales Rafflcsiales Sapindales ASTEKIDAE Solanales Scrophulariales

Lecythidaceae(VAM) 3 Familics Brassicaceae Sapotaceae('?)

THb Hb (Sbj TTHb H b S a THb THb THb THb (Sbj T+ Tr (Sb) Carn Hh Tr Sb

T+ T+ TT+

Protcaceac Sb (Tr) T+ ? Podostcmaceae Aq 2 Families Aq T+ Rhizophoraceae(VAM) Tr+AqSiT+ Para T+ 10 Families(VAM)I 3 Families' Para '? Zygophyllaceae Sb THydrophyllaceae('?j Hb Scrophulariaceae(VAM)Para+ Orobanchaceae I Para Carn Lentibulariaceae

LILIOPSIDA IIIAE ALISMAI Alismatales 3 Families Hydrochant ales Hydrochari taceae Najadales 10 Families COMMELINIDAF. Commelinales Commelinaceae Eriocaulales Eriocaulaceae( ?) Restionales Restionaceae(?) Juncales Juncaceae Cyperales Cyperaceae ZINGIBERIDAE Bromeliales Bromeliaccae'

TTT+ T-

P- EP- EP- EP- EP- EP-EP+EP+ E + P+ P+ PP+

E+ E t EE+

P+ E-

B+ Sp+ B+ S p t B+ Ak+ B+ Sp+ Ak+ Sp+ B+ B+ SP + Aq+ Sp+ Sp+ Gs+Cy+ St+ Tp+ Ak+Cy+

P+ E + Cy+ P+ E + Ak+ P+ E- Cy+Sp+ P-EPPPP-

EEEE-

Sp+AqtGsf Tp+FI+ Ir+ Sp+ Or+ Ir+ Or+ Ir+

Aq Aq Aq

T + P + ET+P+ET+ P t

Hb Aq Hb Hb Hb

Tf

T+ P+ T+ P+ T+ P+

Cy+FI+ Ai+FI+

Epi

T+ P t

EnzSpf

cy + FI +

'?

MYC'ORRHIZAS IN NATURAL ECOSYSTEMS

20 1

ceae and Epacridaceae) almost always form ericoid (or arbutoid) associations (Read. 1983, Reed, 1987), but also form VAM associations in Hawaii (Koske et al., 1990). Mycorrhizas of achlorophyllous plants, including monotropoid plants and some orchids, may superficially resemble ECM but are functionally different (Furman and Trappe, 1971; Harley and Smith, 1983). Root-colonizing fungi that do not form recognized types of mycorrhizal associations can often be observed in natural ecosystems and are especially common in arctic and alpine habitats (Currah and Van Dyk, 1986; Haselwandter and Read, 1982; Kohn and Stasovsky, 1990). Root colonization by one of these fungi (which has dark septate hyphae) can be beneficial to alpine plants (Haselwandter and Read, 1982). Root-colonizing fungi that do not form mycorrhizal associations may benefit plants by conferring disease resistance (Dewan and Sivasithamparam, 1988). These other beneficial rootfungus associations differ from mycorrhizal associations because they show little evidence of morphological or physiological specialization by either organism. The concept that plants have varying degrees of dependence on mycorrhizal associations is gaining acceptance (see Janos, 1980b; Marschner, 1986; Miller, 1987). Detailed examinations of plants in natural ecosystems often show consistently differences between host plants occurring in a particular habitat in the degree (proportion of roots) supporting mycorrhizal associations. These observations have shown that species generally either have consistently high levels of mycorrhizas, intermediate or variable levels of mycorrhizas, or are non-mycorrhizal (Brundrett and Kendrick, 1988; Dominik et al., 1954b; Janos, 1980b; Koske and Gemma, 1990). Plants belonging to these categories have been designated as obligatorily mycorrhizal, facultatively mycorrhizal, or non-mycorrhizal.

Notes to Table 4 (Opposite) '? = mycorrhizal status requires further study: VAM = family also contains species with VAM; 1 = highly reduced roots. H b = herbaceous; Sb = shrub: Tr = tree; Aq = Aquatic herb; Para = parasitic on other plants: Carn = carnivorous: Sa = saline habitats, Epi = epiphytes. T = tannins; P = proanthocyanins; E = ellagic acid; + = present; - = absent; k = present in some. Ak = alkaloids; B = hetalains; Sp = sapniferous; Aq = anthraquinone glycosides; Gs = glucosinolates; St = sterols; Tp = terpinoids: Al = aluminium; Cy = cyanogenic: Ir = iridoid compounds: Or = orobanchin: Enz = proteolytic enzymes.

<'

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M . BRLJNDREIT

Obligatorily mycorrhizal plants have been defined as those which will not survive to reproductive maturity without being associated with mycorrhizal fungi in the soils (or at the fertility levels) of their natural habitats (Janos, 1980b; Kormanik, 1981). These species consistently support mycorrhizal colonization throughout most of their young roots. Faculratively mycorrhizal plants are those that benefit from mycorrhizal associations only in some (of the least fertile) soils in which they naturally occur (Janos, 1980b, see Section 111. E.4). In ecosystem surveys inconsistent mycorrhization (Trappe, 1987) or low levels of mycorrhizal colonization (less than 25%) (Brundrett and Kendrick (1988) have been used to designate facultatively mycorrhizal species when soil fertility levels could not be manipulated. The various factors, including soil nutrient levels, root system characteristics and host plant physiology, that determine the net benefit of mycorrhizal associations to plants are considered in Section 111. E.5. Non-mycorrhizal plants have roots that consistently resist colonization by mycorrhizal fungi, at least when they are young and healthy. These observations provide evidence that intrinsic properties of roots may regulate mycorrhiza formation, at least in some cases and possible mechanisms for this regulation are discussed in Section 111. E.4.c. Nutrient levels and other soil properties (Section 111. E.5) and mycorrhizal propagule dynamics (Section 111. B. 1) can also influence mycorrhizal associations, but in a less consistent manner.

E. Edaphic/Environmental Factors, Plants and Mycorrhizas The events which occur during the life-cycle of a mycorrhizal association are shown in Fig. 2. The formation, survival and dispersal of mycorrhizal propagules have already been considered (Section 1II.B.1). Environmental factors, soil conditions, root properties. rhizosphere effects, etc. that influence the activity of these propagules, mycorrhizal fungus hyphae and roots, as well as the remaining stages in mycorrhizal formation and activity illustrated in Fig. 2 will now be considered.

I . Mycorrhizal Phenology Seasonal variations in the periodicity of root growth and mycorrhizal activity occur in ecosystems and may be substantial enough to change the apparent mycorrhizal status of plants (Allen 1983; Brundrett and Kendrick, 1988; Gay et a l . , 1982; Giovannetti, 1985; Louis and Lim, 1987; Rabatin, 1979). There were no significant seasonal variations in the degree of mycorrhizal colonization of the roots of many herbaceous plants in a deciduous forest community, because only a fraction of their roots were replaced each year (Brundrett and Kendrick, 1988, 1990a).

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These species consequently had high levels of VAM at all times year, but active associations with arbuscules only in their youngest roots, which only comprised a fraction of their root systems. Other species of deciduous forest plants had annual roots with well defined periods of growth and senescence, resulting in abrupt transitions in mycorrhizal colonization levels. These seasonal changes in VAM activity were regulated by root phenology, since VAM associations only formed in young roots and had a limited period of activity (Brundrett and Kendrick, 1990a; Harley and Smith, 1983). Moderate seasonal variations in the extent of mycorrhizas in the roots of European deciduous forest species (Mayr and Godoy, 1989) and salt marsh plants (Van Duin et af., 1989) were also associated with new root production during the growing season. Allen and Allen (1986) found that mycorrhizas delayed the phenology of a grass that received little benefit from the association. Mycorrhizal strategies of plants may be correlated with the environmental conditions prevailing when plants produce new roots, as was observed in a temperate deciduous forest community (see Fig. 3). In this community, most species with root growth in summer (in warm soil) were mycorrhizal, while those with active roots in the spring or fall (in cold soils), had little or no mycorrhizas (Fig. 3). Species in a European

Period of maximum shoot activity

Fig. 3. This chart shows correlations between phenological categories (major period of shoot activity) and mycorrhizal relationships for common species of herbaceous woodland plants in a temperate deciduous forest community (data from Brundrett and Kendrick, 1988). Species with high levels of VAM which grow in the spring o r spring and summer periods (indicated by stars) generally had roots which grow in the summer, even though their shoot activity occurred earlier in the year. (Abbreviations; VAM = vesicular-arbuscular mycorrhial, facultative = low levels of VAM, non-mycorrhizal = without any mycorrhizas.)

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M. RRUNDRETl

grassland community could be divided into similar summer and spring root phenology groups (Fitter, 1986~).Co-existing North American prairie grasses also belong to warm and cool season phenology groups and only those most active at warm times of the year are highly mycorrhizal (Daniels Hetrick and Bloom, 1988; Daniels Hetrick et crl., 1989). Correlations between low ambient temperatures and reduced mycorrhizal colonization have also been reported in a desert shrub (Allen, 1983), cultivated winter wheat (Daniels Hetrick and Bloom, 1983) and with increasing altitude or latitude in arctic and alpine ecosystems (Appendix 1). The predominance of non-mycorrhizal plants in situtions where cold temperatures prevail probably results because these conditions restrict the activity of mycorrhizal fungi (Brundrett and Kendrick, 1988; Daniels Hetrick et al., 1989), but this correlation could also have resulted because cool conditions favoured plants from nonmycorrhizal families (Table 4) for other reasons. Roots forming ECM associations can live for several years (Harley and Smith, 1983). However, ultrastructural studies have shown that only subapical regions of ECM roots have an active host-fungus interface that must be renewed by further root extension (Kottke and Oberwinkler, 1986; Massicotte et al., 1987). Root growth and mycorrhizas formation often parallel shoot development in crop plants, but in many trees root growth is suppressed during periods of rapid leaf formation (Lyr and Hoffmann, 1967) and some herbaceous plants form roots and VAM when shoots are dormant (Brundrett and Kendrick, 1988; Daft et a l . , 1980). Root growth in natural ecosystems generally occurs at times when both temperature and soil moisture conditions are favourable (Allen, 1983; Gregory, 1987; Hayes and Seastead, 1986, Lyr and Hoffmann, 1967; Richards, 1986). Consequently, overall communitywide trends in mycorrhizal activity would be regulated by these environmental constraints on root (and fungus) activity, but variations between hosts would be the result of genetically regulated differences in their root phenology . Experimental investigations have found that the benefits provided by mycorrhizal associations are often reduced by low temperatures and may be eliminated at 5-10°C (Andersen et al., 1987; Chilvers and Daft, 1982; Furlan and Fortin, 1973; Hayman, 1983; Smith and Bowen, 1979). Reduced VAM activity may result because the activity of these fungi stops below a cut-off temperature ( k 7 " C ) (Hayman, 1983) while roots of some plants continue to grow, or because mycorrhizal exchange processes become inefficient at low temperatures. Smith and Bowen (1979) observed that low temperatures reduced VAM initiation, which suggests that fungal activity was affected. There are differences between VAM fungi in the optimum and upper and lower temperature limits for

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spore germination (Tommerup, 1983b; Schenck et a l . , 1975). The activity of ECM fungi is also considerably influenced by temperature (Peredo et al.. 1983; Slankis, 1974). Winter wheat (Dodd and Jeffries, 1986) and Endyrnion non-scriptus (Daft et al., 1980) form VAM associations during mild winter conditions in the UK and some species are mycorrhizal (VAM, ECM or Ericoid) in all but the most extreme arctic and alpine sites (Appendix l), suggesting that some host-endophyte combinations are effective at low temperatures. Under experimental conditions the efficacy of mycorrhizal associations can also be reduced by low light levels (Section III.E.4.a). However, woodland plants growing in heavily shaded conditions can be highly mycorrhizal, but with associations that form and senesce at relatively slow rates (Brundrett and Kendrick, 1990a). Seasonal variations in mycorrhizal spore numbers can occur (Ebbers et al., 1987; Dhillion et al., 1988: Gemma and Koske, 1988b; Gemma et al., 1989; Giovannetti, 1985; Louis and Lim, 1987). In most cases spores are less abundant during periods of mycorrhizal formation and become more numerous during periods of root senescence. These reductions in spore numbers may result from spore germination, limited spore lifespans, or the activities of antagonistic soil organisms-which may coincide with root growth (Mosse and Bowen, 1968; Sutton and Barron, 1972). Peak periods of spore production are generally thought to coincide with periods of fungal resource remobilization from senescing roots (Gemma et al., 1989; Hayman, 1970; Sutton and Barron, 1972). This hypothesis is supported by observations that spore production is greatest when root activity is interrupted by a long dry season or plants are harvested for agricultural purposes (Janos, 1980b; Mosse and Bowen, 1968; Redhead, 1977). However, substantial variations in the timing of spore production occur between VAM fungi associated with a host plant, suggesting that competition between fungi (Section III.B.6) and environmental factors probably also influence spore production (Gemma and Koske, 1989). Spore numbers are not always well correlated with the degree of mycorrhizal formation (Section I11 .Be1) and their germination potential varies at different times of the year (Tommerup, 1983a; Gemma and Koske, 1988b). Other inoculum types are often considered to be more important in natural ecosystems (Section III.B.2), but the total inoculum potential of undisturbed soils been measured (by bioassays) in only a few cases. Bioassays of undisturbed soil did not find large seasonal variations in the inoculum potential of soils in Australia pasture or ecosystem soils (Jasper ef af., 1989c; McGee, 1989; Scheltema et al., 1985a). Fructifications of ECM fungi occur at a specific time of the year (usually in the fall), when spores of these fungi would be abundant in

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their natural habitats. While seasonal variations in the capability of mycorrhizal fungi to initiate associations can occur, mycorrhizal formation requires root growth and factors influencing the latter process appear to be of overriding importance.

2. The Activity of Mycorrhizal Fungus Hyphae in Soils Mycorrhizal fungi form a hyphal network in soil which can obtain and transport nutrients, propagate the association and interconnect plants (Read er al., 1985; Newman, 1988). Production of external hyphae varies between species and isolates of VAM fungi, can be influenced by soil properties and is an important determinant of mutualistic effectiveness (Abbot and Robson, 1985: Graham et al., 1982b; Gueye et al., 1987). Mycorrhizal fungus hyphae are normally thought to obtain poorly mobile nutrients from beyond the zone of nutrient depletion surrounding roots in soils (Section III.E.5), but may also respond to soil heterogeneity. Harvey ef al. (1976) found that most of the ECM roots in a forest soil occurred within organic soil fractions, where litter, woody debris and charcoal decomposition was occurring. Hyphae of these fungi may exploit soil heterogeneity by occupying substrates with lower carbon/nutrient ratios (Coleman et al., 1983). Hyphae of VAM fungi may also preferentially occupy soil organic material (Mosse, 1959; St John ef al., 1983; Warner, 1984), where they produce fine, highly branched, septate hyphae that may have an absorptive function (Mosse, 1959; Nicolson, 1959). Roots also respond to spatial and temporal variations in soil nutrient supply, but they may be less efficient at this than are mycorrhizal hyphae (Section 111.E.5). Mycorrhizal associations can provide the greatest benefit when plants are supplied with forms of phosphorus that dissolve very slowly, so producing highly localized point sources within soils (Bolan et a l . , 1987). Some mycorrhizal fungi apparently can utilize organic or insoluble nutrient sources that are normally thought to be unavailable to plants (Section 1II.F. 1). Absorption of inorganic nutrients by mycorrhizal hyphae and their transport through soil to roots over distances measured in centimetres has been demonstrated by tracers such as 32P (Harley and Smith, 1983; Hayman, 1983). Similar experiments have shown rapid transport of carbon, nitrogen, phosphorus and water by hyphal networks of VAM and ECM fungi (Finlay and Read, 1986; Finlay er al., 1988; Francis et al., 1986; Haystead et al., 1988; Newman, 1988; Read et al., 1985; Ritz and Newman, 1986). In some of these experiments, nutrients transfer occurred between plants of the same or different species. The rapidity of this transfer and its correlation with the presence of mycorrhizal fungi, suggests that it occurs within mycorrhizal fungus hyphae, but this has

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not been fully established (Haystead et al., 1988; Newman, 1988). Francis et al. (1986) reported that mycorrhizal mediated inter-plant nutrient transfer significantly enhanced the growth of recipient plants, but Ritz and Newman (1986) considered the P-transfer rates they measured to be substantially less than uptake rates in the field. Experiments involving competition between mycorrhizal plants suggest that hyphal interconnections provide little benefit to other plants (Section III.F.2). In pot cultures, the proportion of living soil hyphae increases only after root colonization is established and declines rapidly when this process stops (Schubert ef al., 1987; Sylvia, 1988). However, these hyphal growth and viability trends might not occur in ecosystems, where host plants with different periods of root activity could co-operate to support a network of mycorrhizal hyphae and where environmental constraints (temperature, soil moisture levels) would likely be the most important determinants of fungus activity. It has been observed that seedlings form mycorrhizas more rapidly, or to a greater extent when growing near companion plants that are already mycorrhizal (Birch, 1986; Eissenstat and Newman, 1990; Miller et al., 1983; Read et al., 1976). The simultaneous introduction of mycorrhizal fungi and plants to pots in experiments may ultimately impose a greater drain on the host than would occur in nature, because of the costs of creating a new hyphal network from quiescent propagules. The reductions in mycorrhizal colonization resulting from soil disturbance provide further evidence of the importance of pre-existing hyphae as propagules (Evans and Miller, 1988, 1990, Jasper et al., 1989a,b). These hyphal networks may facilitate the absorption and transport of nutrients in soil, since their disruption can reduce the efficacy of mycorrhizal associations in a way that is independent of colonization levels (Evans and Miller, 1990). Mycorrhizal fungus hyphae can influence soil structure by helping to produce humic acids, weathering soil minerals and stabilizing large soil aggregates (Oades, 1984; Perry et al., 1987; Rothwell, 1984), but organic acids and polysaccharides produced by bacteria, fungi and roots and organic debris resulting from root, hyphae, or soil animal activity are also important components of soil structural stability (see reviews by Lynch and Bragg, 1985; Perry et a!., 1987). Major structural contributions to soils by hyphae of VAM or ECM fungi has been observed in arctic communities (Miller, 1982b), sand dunes (Jehne and Thompson, 1981; Rose, 1988), deserts (Went and Stark, 1968), revegetating minesites (Rothwell, 1984) and agricultural fields (Thomas et d., 1986; Tisdall and Oades, 1979). Meyer (1964) considered hyphae of the ECM fungus Cenococcum to be an important structural component of a boreal forest soil and Hunt and Fogel (1983) found this fungus alone can

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comprise as much as 64% of all soil hyphae. Humic acids-organic chemicals which are important components of soil structure and fertility, are produced by partial decomposition of plant residues such as lignin, as well as microbial products-especially fungal melanins (Martin and Haider, 1980; Vaughan and Malcolm, 1985). Similar substances can accumulate in cultures of ECM fungi (Tan et al., 1978). The abundance of mycorrhizal fungus hyphae in many soils suggests that they may be important as source of humic acids as well as influencing soil structural properties.

3. The Rhizosphere and Mycorrhizosphere Whipps and Lynch (1986) describe the rhizosphere as consisting of three zones, the ectorhizosphere (soil in close proximity too roots), the rhizoplane (root surface) and endorhizosphere (apoplastic space within roots). In the rhizosphere soil properties are changed and microbial activity is enhanced by dead cells, mucilages and exudates from roots and these influences are most pronounced near young roots (Curl and Truelove, 1986; Newman, 1985; Uren and Reisenaur, 1988). Root exudates include inorganic ions, sugars, amino acids and organic acids which escape from root cells (Curl and Truelove, 1986). Roots with ECM associations are encased by a mantle of fungal hyphae which would influence or mediate any exudation and cell loss processes that occur, so that the zone or influence surrounding these roots should be called a mycorrhizosphere (Fogel. 1988). The mycorrhizosphere of ECM roots may be expected to have unique properties, including more gradual but sustained exudation, micro-organisms which have evolved to utilize substrates such as trehalose or mannitol, and tolerate defensive secondary metabolites which are of fungal rather than root origin (Fogel, 1988). Fogel (1988) also suggests that mycorrhizal fungi may have a greater influence on rhizosphere properties than host roots, so that the common association of one host with many ECM fungi could result in mycorrhizosphere heterogeneity. It would also be expected that VAM associations (which are a major sink for substrates within roots) would have a substantial influence on mycorrhizosphere properties. Mycorrhizosphere effects, which include many interactions between mycorrhizal fungi and other soil organisms, have been considered in reviews by Fogel (1988), Linderman (1988), Perry et al. (1987) and Rambellini (1973). The VAM mycorrhizosphere may support substantially altered populations of soil bacteria, actinomycetes and fungi when compared with non-mycorrhizal roots (Ames et a l . , 1984; Lawley et a l . , 1982; Meyer and Linderman, 1986b; Secilia and Bagyaraj, 1988). Vancura et al. (1989) observed that soil hyphae of VAM fungi had bacteria growing on

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them and these “hyphosphere” bacteria were a subset of the host rhizosphere population. Soil micro-organisms may enhance (AzconAguilar and Barea, 1985; Meyer and Linderman, 1986a), reduce (Baas et al., 1989a; Daniels Hetrick et al.. 1987; Koide and Li, 1989), or have no effect (Paulitz and Linderman, 1989) on the effectiveness of VAM associations (to increase host growth), relative to pasteurized soil controls. Different soil micro-organisms can enhance or reduce ECM formation, but enhancement is more common (Garbaye and Bowen, 1986, 1989; Richter et al., 1989; Strzelczyk and Kampert, 1987). Garbaye and Bowen (1989) isolated micro-organisms from ECM roots and found that some were capable of enhancing the growth of ECM fungi. They suggest that a community of microbes has evolved in association with ECM roots. However, Summerbell (1989) found that most fungi associated with the mantle of ECM black spruce roots also occurred on non-mycorrhizal roots and roots of a non-host species. Wilkinson et al. (1989) found that some strains of soil bacteria, when co-inoculated with mycorrhizal fungi, enhanced the symbiotic germination of aseptic orchid seeds. Spore dormancy and subsequent activation in response to relatively specific signals, allows fungi to survive in soil when conditions are unfavourable (Sussman, 1976). Germination of VAM spores and subsequent hyphal growth may be enhanced (Azcon, 1987; Azcon-Aguilar ef al., 1986; Azcon-Aguilar and Barea, 1985), or may be inhibited (Paulitz and Linderman, 1989) by the presence of free-living bacteria and fungi isolated from soils. Germination of the spores of some VAM fungi only occurs after the completion of a specific period of dormancy (Bowen, 1987; McGee, 1989; Tommerup, 1983a). Roots exudates can promote germination and hyphal growth from VAM spores (Graham, 1982; Elias and Safir, 1987), within a narrow zone of influence in soils (Smith et al., 1986). Gianinazzi-Pearson et al. (1989) observed that host root exudates and flavonoids they contained elicited rapid germination and hyphal growth responses in a VAM fungus. Volatile factors from roots can also attract soil hyphae or germ tubes from VAM spores (Gemma and Koske, 1988a; Koske, 1982; St John et a l . , 1983). Germination and survival of propagules of VAM fungi (spores) is also influenced by soil moisture, temperature, pH and salinity levels (Daniels Hetrick, 1984; Bowen, 1987). Enhancement of ECM fungus basidiospore germination occurs in the presence of some yeasts and filamentous fungi, but host root exudates often produce a greater response (Ali and Jackson, 1988; Fries, 1987a; Theodorou and Bowen, 1987). Hyphae of ECM fungi can be attracted by exudates of both host and non-host species (Duddridge, 1987). The rhizosphere influence is most pronounced near young roots (Curl

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and Truelove, 1986; Newman, 1985; Uren and Reisenaur, 1988), which may explain their greater attraction to mycorrhizal hyphae (Chilvers and Gust, 1982; Mosse and Hepper, 1975; Smith and Walker, 1981). Mycorrhizal hyphae require several days to respond to the presence of a growing root tip, so would first interact with root cells that are at least 2 days old and take several more days to establish an effective association with a Hartig net or arbuscules (Alexander et al., 1989; Piche and Peterson, 1985). Gemma and Koske (1988a) found that host root growth and lateral root production could be stimulated by the presence of VAM fungi before an association was formed. Plants belonging to certain families have mutualistic associations with N2-fixing bacteria contained within root nodules (Gibson and Jordan, 1983; Torrey, 1978) and associative N2-fixing bacteria may occur in the rhizosphere of other plants (Giller and Day, 1985; Ho, 1988; Pacovsky, 1989), or within ECM roots or basidiocarps (Li and Hung, 1987). There can be interactions between mutualistic nitrogen-fixing bacterial associations and mycorrhizas, which may result because plants with dual associations tend to have higher phosphate requirements (Barea and Azcon-Aguilar. 1983). Soil bacteria which solubilize rock phosphate may increase nutrient uptake by mycorrhizal associations (Piccini and Azcon, 1987). Additional mycorrhizosphere interactions that may occur include inhibition of pathogenic organisms (Section III.E.6) and competition with other soil organisms for nutrients (Section 1II.F.1). The rhizosphere consists of two zones inside and outside the root (the endorhizosphere and ectorhizosphere) which are considered to be more or less contiguous (Whipps and Lynch, 1986). However, roots often have an exodermis with suberin lamellae and Casparian bands which function as a peripheral apoplastic permeability barrier (Peterson, 1988). It seems that many roots which contain VAM associations also have an exodermis- the role of which should be considered in mycorrhizal studies (Brundrett and Kendrick, 1988, 1990b; Smith et a!., 1989). The exodermis is a peripheral root layer, which is similar in structure and apparently also in function to the endodermis, that can reduce root permeability to water and mineral nutrients (see reviews by Drew, 1987; Passioura, 1988; Peterson, 1988) and thus may restrict the diffusion of exudates from roots. However, these substances would still be available in the apoplatic spaces within the root and this internal-root exudation could help maintain quiescent VAM fungi perennating within roots (Section III.B.1). Cell walls in the exodermis and fungal mantle hyphae can function as permeability barriers which would isolate the site of nutrient exchange to Hartig net hyphae and adjacent epidermal cells in angiosperm roots with ECM associations (Ashford et a l . , 1989). Endodermal and exodermal Casparian bands may also help to prevent

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losses during nutrient-exchange processes in roots with VAM associations. It has been suggested that mycorrhizas are more important to plants with a suberized exodermis than those without this layer (Brundrett and Kendrick, 1988; Von Endrigkeit, 1933). Unfortunately, little is known about root exudation in natural ecosystems, or the influence of mycorrhizal associations or root anatomy features such as the exodermis on this process.

4. Regulation of Mycorrhizal Associations Thousands of unsuccessful root-fungus contacts may be required for every successful establishment of a pathogenic fungus within roots (Huisman, 1982), but unsuccessful contacts between mycorrhizal fungi and host roots apparently are rare. However, there are some plants in natural ecosystems which unyieldingly resist the advances of mycorrhizal fungi by as yet unexplained mechanisms (Section III.D.2). The roots of these non-mycorrhizal species may co-exist with mycorrhizal roots of other species, so the problem is not always a lack of inoculum (Brundrett and Kendrick, 1988). The failure of mycorrhizal fungi to colonize roots could occur at a number of the stages illustrated in Fig. 2, but most likely involve a lack of hyphal attraction to the root, or prevention of hyphal penetration of the root, since hyphal activity is aborted before colonies form inside the root. It has been proposed that mycorrhizal fungus ingress can be prevented by physical or chemical barriers or the absence of a factor which promotes hyphal growth (Bowen, 1987; Brundrett and Kendrick, 1988; Testier et a f . , 1987). Chemical, morphological and physiological properties of roots that may influence mycorrhizal formation are considered below. (a) Root morphology and rnycorrhiza formation. Associations between soil fungi and roots are thought to involve contact between fungus propagules with limited mobility and more rapidly growing roots; the number of these contacts which are successful will ultimately determine the intensity of the association (Huisman, 1982). If we assume that most root-fungus contacts will result in root colonization, the intensity of a mycorrhizal association will be regulated by an interaction between (i) the distribution of active propagules (hyphae or germ tubes) in the soil and the rapidity of their response to the presence of roots and (ii) the number of growing tips, growth rate, spatial distribution and duration of susceptibility of host roots (mycorrhizal initiation occurs in young roots and effective associations have limited lifespan) (Brundrett and Kendrick, 1990a; Hepper, 1985). We would normally only expect mycorrhizal inoculum to be in short supply in disturbed sites (Sections III.B.4,

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III.F.4), or areas where extreme climatic or edaphic conditions favour plants with low levels of mycorrhizas (Section III.F.5). The balance between fungus and root activity would result in relatively low levels of mycorrhizal formation if plants have active root systems in which highly branched, rapidly growing, narrow and relatively short-lived roots predominate. As will be considered further in the next section, species with low mycorrhizal dependency often have root systems of this type. The narrow cortex of fine roots may also limit mycorrhization because relatively few exchange sites (arbuscules) would result from each contact with a mycorrhizal fungus, whose hyphae may reach less of these rapidly growing roots while they are still young enough to form an association. However, species with extensive root systems may require lower levels of mycorrhizal colonization (measured as a proportion of root length) to achieve similar nutrient inflow rates than species with coarse roots. This hypothesis should be examined by comparing phosphorus inflow rates, the total volume of mycorrhizal colonies, or numbers of arbuscules between roots of facultatively and obligately mycorrhizal species. Host roots are capable of much faster growth than ECM fungus hyphae, which also must contact young root cells to form an association, so formation of ECM associations apparently only occurs in roots with reduced rates of elongation (Bowen and Theodorou, 1973; Chilvers and Gust. 1982; Duddridge. 1987). This may explain why only plants belonging to a limited number of families form ECM associations and members of these diverse families have independently evolved similar heterorhizic root systems in which some lateral roots have very limited elongation (Brundrett et a[., 1990; Kubikova, 1967). The interdependence of root growth and fungus activity should be considered to be one of the main attributes of mycorrhizal associations. Some idea of the relative contribution of host and endophyte features to regulation of VAM formation can be obtained by contrasting different combinations of host plants and mycorrhizal fungi, Consistent features of mycorrhizal morphology that are associated with particular mycorrhizal fungi have sometimes been used to identify endophytes (Abbott, 1982: Agerer, 1986; Haug and Oberwinkler, 1987), but root features can also regulate mycorrhizal morphology. Their is evidence that arbuscule formation is primarily under fungal control but there were some differences in the size of arbuscules that may be attributed to the host (Alexander ef al., 1989; Lackie et al., 1987). Gallaud (1905) observed that two distinct VAM morphology types occurred in the roots of different species, (i) the Arum series where hyphae proliferated by growing between cortex cells and (ii) the Paris series where hyphae formed coils within cells. These morphological distinctions arise because hyphae grow along longitudinally continuous intercellular air spaces in

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Arum series hosts, but these longitudinal channels are absent in Paris series roots (Brundrett and Kendrick, 1988, 1990a,b). This provides morphological evidence that the efficiency of VAM associations (arbuscule location and abundance within roots) can be regulated by root anatomy features and it seem likely that some aspects of root form evolved as means to control these associations. The subepidermal layer of roots of many angiosperms differentiates into a suberized exodermis with Casparian bands and suberin lamellae (Brundrett and Kendrick, 1988; Peterson, 1988; Shishkoff, 1987). Cells in this layer can play an important role in the resistance of roots to plant pathogens (Brammall and Higgins, 1988) and may also restrict the passage of mycorrhizal fungi. In roots with this potential barrier, it may consist of uniformly suberized cells or alternating suberized long cells and unsuberized short cells (Shishkoff, 1987). In plants with short cells mycorrhizal fungus penetration typically occurs through these “passage cells” (Bonfante-Fasolo and Wan, 1989; Brundrett and Kendrick, 1988; Gallaud, 1905), while in other cases fungal entry has been observed to proceed uniform exodermis suberization (Brundrett and Kendrick, 1990a). It is not known if cells with suberized walls restrict the passage of mycorrhizal fungi (which may lack the enzymes necessary to degrade them) or if these fungi avoid them by following paths of lesser resistance into roots. The Hartig net of ECM associations is normally confined to the epidermis of the roots of angiosperm hosts (Alexander and Hogberg, 1986; Massicotte et al., 1987, 1989), but occupies much of the cortex of gymnosperm roots (Harley and Smith, 1983; Kottke and Oberwinkler, 1986). It seems likely that structural or chemical properties of outermost cortex cells prevent further ingress by ECM fungi in angiosperm roots. Inner-cortex cells with thick, highly refringent walls also function as fungal barriers in the roots of some gymnosperms (Brundrett el al., 1990). The distribution of ECM fungus hyphae can be correlated with the distribution of “pectins” (acid polysaccharides that are relatively flexible) in root cell walls (Nylund, 1987). Plants with mycorrhizal associations become less important than non-mycorrhizal species in some habitats with adverse soil conditions (Appendix 1) and this may be correlated with root structural specializations of plants in these communities (Section 1II.C. 1). Mycorrhizal associations generally are sparse or absent in the roots of hydrophytes which generally have adapted to growth in anoxic substrates by having large cortex air spaces in their roots. We would expect these roots to be structurally less compatible with VAM associations, because mycorrhizal fungus hyphae would impede oxygen flow if they grew along air channels and in the case of roots with large spaces, there would be few

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remaining cells with which to form an association. Plants which do not have roots as adults (including some parasitic, epiphytic, or submerged aquatic members of families in Table 4) would not be able to form mycorrhizal associations. Plants adapted to growth in regions were soils are dry most of the time, often have deep root systems or shallow roots that proliferate rapidly when the soil becomes wet (Eissenstat and Caldwell, 1988; Franco and Nobel, 1990; Richards and Caldwell, 1987). The rapid growth of these roots and the brief period of their activity could make them inefficient mycorrhiza formation. (b) Host plant physiology. Cost benefit analysis can be used to weight the benefits provided by mycorrhizal associations (enhanced mineral nutrient uptake) against the costs (carbon supplied by the host) of the association (Section III.E.7). In situations where the cost of mycorrhizas outweigh their benefits, one would expect the host plant to restrict mycorrhizal formation in some way. Situations where mycorrhizal formation is not regulated by a balance between root growth and mycorrhizal fungus propagule activity (as was considered above) are most likely to occur when the host already has ample supply of nutrients or is unable to supply sufficient amounts of carbohydrates to the fungus. The inability to supply carbohydrates may explain why plants growing at suboptimal light levels and/or at low temperatures (both of which reduce photosynthesis) can have reduced or inefficient mycorrhizal associations (Graham el al., 1982a; Hayman, 1983; Son and Smith, 1988). It has also been proposed that the formation of ECM associations is regulated by root carbohydrate levels (which are influenced by plant mineral nutrition and light), although plant growth hormones (auxins or ethylene) produced by mycorrhizal fungi may also be involved (Harley and Smith, 1983; Nylund, 1988; Rupp et a l . , 1989). It has often been observed that VAM formation can be considerably reduced by growing plants at high phosphorus levels, even in the presence of abundant mycorrhizal fungus propagules (Amijee et a l . , 1989; Menge et a [ . , 1978; Mosse, 1973; Thomson et a l . , 1986). although at much lower levels phosphorus addition can enhance VAM formation (Bolan et a l . , 1984; Koide and Li, 1990). The greatest reduction in mycorrhizal fungus activity apparently occurs because of phosphorus levels and processes inside the root (Jasper et a l . , 1979; Koide and Li, 1990; Menge et al., 1978; Thomson et al., 1986). Son and Smith (1988) reported that reductions in VAM caused by high phosphorus were more severe at low light levels and Thomson et a/. (1986) associated high phosphorus level VAM inhibition with reductions in soluble carbohydrates within roots and their exudates. High phosphorus levels, which are correlated with reductions in root carbohydrate levels, can also

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21s

inhibit ECM formation (Marx et al., 1977). Phosphate deficiency can increase root exudation by a potential host plant, which may be correlated with the degree of VAM formation (Elias and Safir, 1987: Graham et al., 1982a: Thomson et a l . , 1986). However, in other cases their appears to be no clear relationship between root exudation and VAM formation (Ocampo and Azcon. 1985; Schwab et al., 1983). Mosse (1973) and Amijee et al. (1989) observed that many VAM fungus entry points aborted when Alliurn species where grown at very high phosphorus levels. These reports of reduced mycorrhizal formation caused by low light levels or high phosphorus supply in experiments may be applicable to some agricultural situations but are less relevant to plants in nature where such high phosphorus levels are unlikely to occur and plants generally are exposed to the fertility and light levels to which they are adapted.

(c) Chemical root features. It has frequently been observed that plants belonging to families such as the Brassicaceae (Hirrel et al., 1978; Medve, 1983), Cyperaceae and Juncaceae (Powell, 1975), Proteaceae (Lamont, 1982) or Zygophyllaceae (Khan, 1974, Trappe, 1981) rarely form mycorrhizas. Additional families and orders of plants which are predominantly non-mycorrhizal are listed in Table 4, which is compiled from many sources. Some of these families contain species, which are aquatics, epiphytes, or parasites with much reduced roots, but in other cases their lack of mycorrhizas cannot be explained by habitat preferences. Major chemical constituents of these families (after Cronquist, 1981) are also listed in Table 4. Many of the secondary metabolites listed in Table 4 can also be found in plants from predominantly mycorrhizal families. However, it would seem that plants in “non-mycorrhizal families” are more likely to accumulate chemicals which are considered to be evolutionarily advanced (see Cronquist, 1977, 1981) in comparison with more primitive chemical components, such as phenolics, which are often scarce or absent in these same families. Plants have evolved a very wide diversity of secondary metabolites, but the function of these chemical is often unknown. Potential roles include interactions with other plants, herbivores and pathogens (Bell, 1981). It seems likely that many of these chemicals are accumulated in plants for defensive purposes and that relatively advanced chemicals evolved because herbivores and pathogens developed resistance to older chemicals (Cronquist, 1977, 1981). There is reason to believe that many of the secondary metabolites accumulated by plants belonging to the families listed in Table 4 (cyanogenic glucosides, betalains, alkaloids, etc.), can be antagonistic to fungi. It has been suggested that these

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chemicals are responsible for the absence of niycorrhizas in plants i n which they accumulate (Bowen, 1987; Hayman et al., 1975; Iqbal and Qureshi, 1976: Kumar and Mahadevan, 1984: Lesica and Antibus, 1986) and there is some evidence to support this hypothesis. Flavonoids include a diverse group of common root constituents, some of which have been implicated in plant protection from fungal pathogens, insects and nematodes, as well as allelopathic interactions (Rao, 1990). Morandi et al. (1984) reported that VAM colonization enhanced flavonoid production in soybean roots and in a later study (Morandi, 1989) could find no immediate correlations between their concentration (as influenced by pesticide treatments) and VAM colonization. Gianinazzi-Pearson et al. (1989) observed that flavonoids in root exudates produced by soybean roots induced rapid spore germination and hyphal growth responses in a VAM fungus, but exudates from Lupinus roots did not. Members of the genus Lupinus (in the predominantly VAM family Fabaceae) are resistant to colonization by VAM fungi (Trinick, 1977) and can inhibit mycorrhizal formation in adjacent roots of other plants (Morley and Mosse, 1976). Lupinus roots are known to contain isoflavonoids and alkaloids with antifungal properties (Lane et al., 1987; Wink, 1987), which may have an adverse influence on VAM fungi if contained in their exudates. Betalains replace flavonoids in most members of the predominantly non-mycorrhizal order Caryophyllales (Cronquist, 1981) and their influence on fungi warrants investigation. Glucosinolates (isothiocyanate derivatives), which are characteristic components of the Brassicaceae and other Capparales (Cronquist, 1981; Testier et a f . , 1987; Table 3), also have antifungal properties (Drobnica et a f . , 1967). However, Testier er al. (1987) found that these compounds are also contained within many mycorrhizal plants. In a study of Canadian deciduous forest plants most were found to have VAM associations, but several species had roots that remained free of mycorrhizal fungi (Brundrett and Kendrick, 1988). These include Chelidonium rnajus and several other members of the families Papaveraceae and Fumariaceae, which are known to accumulate fungistatic isoquinoline alkaloids (Gheorghiu er a f . , 1971; Hakim et al.. 1961; Hejtmankova e f al., 1984). Of these related plants, C. rnajus and Dicentra spp. were non-mycorrhizal, while Sanguinaria canadensis had VAM in its fine, but not its coarse lateral roots. This restriction of VAM to fine S . canadensis roots was negatively correlated with the occurrence of orange substances (that were likely to be isoquinoline alkaloids) in coarse roots. Kumar and Mahadevan (1984) found that all of the Indian medicinal plants they examined (which contain large amounts of secondary metabolites) were non-mycorrhizal. The above

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evidence suggests that mycorrhizal strategies can be correlated with plant chemistry, but there is insufficient information is to test this hypothesis for many of the families listed in Table 4. Tommerup (1984) compared the time-course of VAM colonization in Brassica and Trifolium roots. She found that Brassica roots induced considerably lower rates of spore germination, germ tube growth and successful appressoria formation than occurred in the presence of Trifoliurn roots (a potential host). In addition, few arbuscules and no vesicles were produced in the Brassica roots. Tommerup (1984) suggested that volatile or soluble compounds in Brassica roots were probably responsible for the observed inhibition, but a shortage of specific substances required by the fungus could also have been a factor. El-Atrach et al. (1989) found that roots of the non-host Brassica oleracea reduced mycorrhiza formation in Medicago sativa when spores, but not when bulk soil was used as inoculum. They also observed that Glornus mosseae spore germination was not influenced by Brassica root exudates, but was inhibited by a volatile factor produced by these roots. These experiments provide some evidence that secondary metabolites produced by non-host plants can adversely influence mycorrhizal fungi. However, Glenn ef al. (1988) found that VAM formation was not correlated with the glucosinolate content of Brassica cultivars. In their study, host roots stimulated VAM fungus growth, but roots of Brassica did not, which may suggest that they were lacking a diffusible hyphal growth promoter. Ocampo et al. (1980) and Testier ef af. (1987) have also proposed that the absence of mycorrhiza formation in non-mycorrhizal plants was the result of internal root or cell wall properties rather than secondary metabolites, but there is little evidence to support this theory. Further evidence that secondary metabolites can influence mycorrhizal fungi is provided by reports that non-mycorrhizal plants and their residues sometimes have an allelopathic influence on mycorrhizal formation (Section III.F.3.a). The potential role of secondary metabolites in regulating mycorrhizal relationships is complex and would involve many factors that are not well understood. These factors include (i) the accumulation and compartmentalization of secondary metabolites in active or inactive forms, (ii) the susceptibility of mycorrhizal fungi to these chemicals, (iii) the quantity and activity of these chemicals released as soluble o r volatile forms from roots and (iv) if they are released or activated by plants in response to mycorrhizal fungi. Many secondary metabolites are toxic to the cells that make them, so are stored in inactive forms, are compartmentalized in vacuoles, or are detoxified in the cytoplasm (Matile, 1987; McKey, 1979). Thus, it may be possible for mycorrhizas to form an association with roots without

2 18

M . HKLINDKETT

being exposed to appreciable quantities of chemicals they contain. Suberized or lignified structures (which may function as constitutive mechanical defences) were usually well developed in the roots of obligately mycorrhizal deciduous forest plants, but were conspicuously absent from those of non-mycorrhizal species (Brundrett and Kendrick, 1988). This may provide evidence that other more potent chemical defences are present. Primitive chemical constituents of plants, frequently include phenolics and flavonoids which are also potentially fungitoxic (Friend, 1981; Rao, 1990). However, most roots contain phenolics, so mycorrhizal fungi may well have developed tolerance to these chemicals or detoxification mechanisms (Duchesne et al., 1987), in a similar fashion to non-biotrophic plant pathogens (De Wit, 1987) and may even use them as signals to indicate susceptible roots (Brundrett and Kendrick, 1990b; Gianinazzi-Pearson ef a l . , 1989). Wacker et al. (1990) observed that ferulic acid (a simple phenolic compound) produced by asparagus roots was inhibitory to VAM fungus growth at higher soil concentrations. Unlike physical defences, chemical defences apparently decline in effectiveness as roots age. Low levels of colonization have been reported in the roots of non-mycorrhizal plants, especially when they are old or growing in close proximity to mycorrhizal roots (Glenn et al., 1985; Hirrel et a l . , 1978; Miller et al., 1983; Ocampo et al., 1980; Ocampo and Hayman, 1981). These atypical VAM infections did not supply nutrients to non-mycorrhizal plants (Ocampo, 1986) and often occur in moribund roots (Ocampo and Hayman, 1981: Brundrett and Kendrick, 1988). It is possible that many of the reports of VAM in non-mycorrhizal families (see Harley and Harley, 1987; Testier ef al., 1987; Table 2) represent non-functional infections (without arbuscules) that were induced by the presence of host roots or occurred in senescent roots. Non-mycorrhizal plants may passively release secondary metabolites into their rhizosphere. Morphological comparisons between the roots of non-mycorrhizal and highly mycorrhizal plants provide evidence (such as highly active and extensive roots systems and the absence of an exodermis) that the roots of non-mycorrhizal plants would be likely to produce more exudates (Table 5 ) . Darnall and Burns (1987) were able to detect glucosinolates (but not free HCN) released into the rhizosphere by plants, but the influence the concentrations they measured would have on fungi is unknown. The volatile factor released by Brassica roots that El-Atrach et al. (1989) found could inhibit Glomus spore germination may have resulted from glucosinolate breakdown. Rhizosphere factors have been implicated in the inhibition (in the case on non-mycorrhizal plants) or promotion (in the case of potential hosts) of mycorrhizal fungus activity. It is not known if mycorrhizal fungi can

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

219

Table 5 Root system features correlated with mycorrhizal dependency Typical trends in root features

Mycorrhizal dependency continuum High

Low

Surface area of absorbing rootso root length/biomass ratioh lateral root branching orders branching frequency root hairs

Low Low Few Sparse Fewlshort

High High More Frequent Manybong

Root system activity root growth responsivenessc

Low Slow Little

High Fast High

Root lifespan (in primary growth)

Monthslyears

Weeks/months

Protective features structurald chemical'

Well developed Relatively primative

Weakly developed Relatively advanced

Rhizosphere influencesf

Slight

May occur

Root activity at low temperatures

Usually stops

May be considerable

Formation of mycorrhizas

Efficient Well regulated

Inefficient May be inhibited

" Relative to plant biomass; specific root length; roots respond to temporary or localized soil conditions; suberization or lignification of primary root structures; accumulated secondary metabolites may be relatively primative (tannins etc.) or advanced (alkaloids, cyanogens etc.): f that influence the availability of soil nutrients.

detect differences in the rhizosphere properties of susceptible vs nonsusceptible roots (which would result in fungus gene expression changes) or if they are influenced by these properties in other ways (perhaps through physiological processes). A distinction also needs to be made between chemicals which are part of a plant's constitutive defences and those induced as a result of injury. Codignola et al. (1989) did not detect increased cell-wall phenolic production in Allium epidermal and exodermal cells after colonization by a VAM fungus, but phenolic wall depositions in these cell layers can be an important part of a resistant host plant's response to invasion by a pathogenic fungus (Brammall and Higgins, 1988). Amijee et al. (1989) and Mosse (1973) observed that very high phosphorus levels resulted in aborted VAM fungus penetration of roots by a reaction at the root periphery and their illustrations suggest that this process occurred within exodermal short cells. Allen et al. (1989a) found that a similar wounding response resulted in aborted VAM fungus colonization of roots of the

220

M. BRUNDRETT

non-host plant Salsola kali. Duc er al. (1989) produced mutants of two normally mycorrhizal species which would not allow VAM fungi to occupy their roots. Their results suggest that host plants have potential mechanisms for excluding fungi from their roots which can be used to exclude pathogens, but which are normally not activated during mycorrhizal formation. Interactions between biotrophic plant pathogens and compatible hosts apparently also involve a failure of the host to respond adequately to the presence of the fungus (De Wit, 1987). Roots of host plants forming VAM or ECM associations often intermingle in natural ecosystems and would be exposed to propagules of both types of fungi, but normally only form one type of association (Section III.D.2). This may suggest that roots of ECM hosts repel or fail to attract VAM fungi and vice versa. It would appear that the major difference between facultatively mycorrhizal and non-mycorrhizal plant species is that facultative species lack the ability possessed by non-mycorrhizal species to exclude mycorrhizal fungi from their roots completely. Plants in both these categories are more likely to grow in habitats where mycorrhizal fungi would be of limited benefit (Section III.F.5). As a consequence, it may have been advantageous for these plants to evolve mechanisms which prevent mycorrhizal colonization and the resultant loss of energy to the fungus. Alternatively, they might have been excluded indirectly by a process evolved as a defence against pathogenic organisms, if mycorrhizas no longer were of much value. It appears likely that non-mycorrhizal plants have evolved potent mechanisms for excluding mycorrhizal fungi from their roots, which may include the stockpiling of fungistatic chemicals such as the alkaloids in Chelidonium roots, active defensive reactions, the absence of signals mycorrhizal fungi use to “recognize” susceptible roots, or other forms of chemical or structural incompatibility. There is insufficient evidence to determine which of these hypothetical mechanism(s) are used by non-mycorrhizal plants in natural ecosystems. These complex interactions provide evidence of the co-evolution of host plant chemical root properties and mycorrhizal fungi.

5 . Mycorrhizal Dependency (a) Soil properties. Plants in natural ecosystems have varying degrees of dependence on mycorrhizal associations which are the result of inherent properties of the plants themselves and the availability of nutrients in the soils in which they naturally occur (Janos, 1980b). Mycorrhizal dependency is simply a measure of the benefit provided by mycorrhizas and will depend on the relative contribution of root and

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

22 1

mycorrhizal mediated nutrient uptake to plants. The outcome of this competition between the roots themselves and their mycorrhizal associations would depend on the root system properties considered in the next section and soil characteristics influencing nutrient availability. The supply of a particular mineral nutrient to a plant depends on its availability and mobility in soils, as well as the plants internal requirement for that nutrient (Marschner, 1986: Russell, 1977). Phosphorus is generally considered to be the most important plant-growth limiting factor which can be supplied by mycorrhizal associations, because of the many abiotic and biotic factors which can restrict its mobility in soils (Harley and Smith, 1983: Hayman, 1983: Marschner, 1986; Wood et af., 1984). Reductions in the benefit provided by mycorrhizal associations (mycorrhizal dependency) to plants caused by increasing soil phosphorus levels have often been observed (Baylis, 1975: Crush, 1973b; Daniels Hetrick er af., 1989, 1990: Gerschefske Kitt et af., 1988; Johnson, 1976). In general, both mycorrhizal and non-mycorrhizal roots access the same “available” sources of soil phosphorus (Harley and Smith, 1983; Hayman, 1983), but there is some evidence that mycorrhizas may have a greater effect when phosphorus is present in less soluble forms. Bolan et al. (1987) observed that VAM provided the greatest benefit when plants were supplied with iron phosphates, which are poorly soluble, so would provide highly localized phosphorus sources. Jayachandran et af. (1989) observed enhanced growth of plants with VAM (but not non-mycorrhizal plants) after the addition of an iron chelator to an infertile soil. These studies suggest that a high phosphorus fixing capacity within a soil may contribute to mycorrhizal dependency by impeding the uptake of “available phosphorus” to plants. Sainz and Arines (1988) reported that soil sterilization and the activities of mycorrhizal fungi can result in changes in the relative proportion of phosphorus contained in different organic or inorganic soil fractions (as measured by different extraction procedures). The abundance of organic vs inorganic nutrient sources (see Section F. 1) and variations in soil phosphorus-fixing capabilities, may influence the relative availability of phosphorus to roots and mycorrhizal fungi in natural soils, but this requires further investigation.

(b) Root properties. Considerable variations in root system extensiveness, geometry, depth distribution and plasticity occur between plant species (Crick and Grime, 1987; Fitter, 1987; Richards and Caldwell, 1987). Plants with extensive (highly branched, fine, long roots with numerous root hairs) have often been observed to have sparse mycorrhizas in natural ecosystem soils or derive little benefit from mycorrhizas in experiments. Observations of this type have involved plants in Amazonian rainforests (St John, 1980b), New Zealand forests (Baylis,

222

M . BRUNDRETT

1975; Johnson, 1976), prairies (Daniels Hetrick et al., 1988; Miller, 1987), temperate deciduous forests (Fig. 4), arctic communities (Miller, 1982a) and peatlands (Hoveler, 1892). Mineral nutrients such as phosphorus have very limited mobility in soils so that depletion zones, where much of the available nutrient has been utilized, quickly form around roots (Kraus et al., 1987; Marschner, 1986; Russell, 1977). Thus to obtain more phosphorus, plants must bypass the depletion zones surrounding existing roots by further root activity elsewhere in the soil. The outcome of this quest for phosphorus (and other relatively immobile soil resources) should largely be determined by the surface area of a plant's root system, which in turn is a product of the length and diameter of roots, the abundance and length of root hairs and branching properties (architecture). These properties have been found to be positively correlated with a plant's ability to absorb phosphorus in a variety of situations (Bolan ef al., 1987; Cardus. 1980; Crush, 1974; Fohse e f al., 1988; Itoh and Barber, 1983, etc.). Mengel and Steffens (1985) found that absorption of potassium (a nutrient of intermediate soil mobility) was also correlated with root length differences between species. Some plants, like rye grass, have narrow. highly branched roots with

100

80

-

60

-

40

-.

'

c

0

0

K

0

o 0

AA

Trees

0

Spring

0

0 Spring-Summer

. A , A

0 -r 5

0

an

.

2o

0

U0A A

A

Summer

0

Fall 1

I

10

15

A 0

m m p 20

25

Active area Index

Fig. 4. The relationship between active area index (which is an arbitrary sum of root diameter, branching order, hair developments, lifespans and suberization indices) and VAM colonization (% of primary root length). Data was obtained from a study of deciduous forest plants (Brundrett and Kendrick, 1988) and includes common trees, shrubs and herbaceous plants-which have been separated into the phenological categories used in Fig. 3.

MYC'OKKHIZAS IN NATURAL ECOSYSTEMS

223

numerous root hairs forming a vast surface area in contact with the soil: one rye plant can produce 600 km roots with a 600m' surface area (Dittmer, 1937). Other plants, such as onion (Bhat and Nye, 1074) and most of the temperate forest plants examined by Brundrett and Kendrick (l988), have coarse, infrequently branched roots with few root hairs, resulting in substantially less soil contact. Root hairs can make a large contribution to root surface area, but usually are short-lived structures and considerable variations in their length and abundance occur between species (Dittmer, 1949). or as a result of soil conditions (MacKay and Barber, 1985). In ecosystem studies, total biomass is the parameter most often used to quantify roots, but provides much less information than root length or specific root length (length/unit root weight) data (Fitter and Hay, 1987: Kummerow, 1983). Parameters that could be used to predict nutrient absorption are; total root biomass < fine root biomass < root length < root surface area < root surface area and activity < rhizosphere volume, arranged in increasing order of predictive ability. These measurements would all be relative to whole plant (or shoot) biomass. It is sometimes possible to provide good estimates of root surface area from root biomass after careful study of roots in a particular ecosystem (Kummerow, 1983). Investigations of crop plants have shown that the biomass, length, volume, growth rate, activity, depth and water conductance and strength of roots can influence cultivar responses to nutrient levels, temperature, drought, hard soil layers, wind, disease and poor drainage (O'Toole and Bland, 1987). There are considerable variations between plant species in the degree of root system architecture (distribution and branching pattern) changes that occur in response to localized sources of nutrients and water in soil, but it appears that variations occurring within species are considerably less than variations in root system plasticity between species (Crick and Grime, 1987: Drew and Saker, 1975, 1978; Fitter, 1987; Grime er a f . , 1986; St John er al., 1983). The responsiveness of plant root systems to small scale or short duration changes in water or nutrient availability is thought to be an important determinant of their success during competition for soil resources (Campbell and Grime, 1989; Franco and Nobel, 1990; Jackson and Caldwell, 1989). Powell (1974) found that non-mycorrhizal Carex coriacea plants grown at low soil phosphorus levels had a greater specific root length (finer roots) but a similar root:shoot ratio than plants grown at a higher fertility levels. Moderate reductions or increases in the relative area of root systems have been observed when mycorrhizal plants were compared to non-mycorrhizal controls (Anderson and Liberta, 1989; Daniels Hetrick et al., 1988; Graham, 1987; Price er al., 1989). However, these responses (which probably result from

224

M . I3KllNI>KFTT

competition for carbohydrates or differences in phosphorus nutrition) are apparently much less plastic than those of plants with little mycorrhizal dependency. Koide el al. (1988) observed that wild oats had a root system that is more responsive to soil nutrient levels than cultivated oats and the latter received more benefit from mycorrhizas. Root system plasticity (or responsiveness to changes in soil heterogeneity) apparently allows plants to forage for soil nutrients without relying on mycorrhizal associations, but probably results in high root system maintenance costs (due to growth and respiration rates). Extensive, highly active root systems alone may not be enough to ensure adequate mineral nutrient capture for non-mycorrhizal plants. Fohse er al. (1988) observed that onion, tomato and bean plants (which often benefit from VAM) had substantially lower phosphorus uptake efficiency per unit root length than rape and spinach (non-mycorrhizal species). Van Ray and Van Diest (1979) found that buckwheat (Fagopyrum escdentum), a plant that has been reported to be non-mycorrhizal (Harley and Harley, 1987), was able to obtain phosphorus from sources that were unavailable to other species. Some plants have the ability to change rhizosphere conditions, such as pH, which influence nutrient availability (see Marschner, 1986; Uren and Reisenaur, 1988). However, this influence of roots on soil will also depend on the extensiveness and activity of their root system, since young roots are the primary source of exudates (Curl and Truelove, 1986; Uren and Reisenaur, 1988). Mycorrhizal fungus hyphae primarily function by effectively increasing the soil volume from which immobile nutrients are absorbed and provided to roots (Hayman, 1983; Harley and Smith, 1983). By comparing their relative diameters, Harley (1989) has estimated that, per unit length, fungus hyphae would be approximately 100 times less expensive to form and maintain than roots. Mycorrhizal associations and extensive, highly active root systems are two alternatives in the quest for nonmobile soil nutrients and mycorrhizal fungus hyphae should be a more cost-effective means of exploring large soil volumes. It is becoming increasingly apparent that variations between species in roots system properties such as extensiveness, responsiveness, growth rates, rhizosphere modification and mycorrhizal colonization (which are often correlated) correspond to substantial differences in their below-ground strategies for nutrient uptake. Root system features that are often correlated with the mycorrhizal dependency of plants are summarized in Table 5. The root system properties described above can be used to formulate an active area index, which is the sum of separate arbitrary scales referring to root hair length and abundance, root diameter, branching

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

225

orders, root suberization and root lifespan data. An index of this kind has been formulated for 26 temperate deciduous forest plants and contrasted with their degree of mycorrhizal colonization in Fig. 4. The inverse correlation between this index (which provides a crude estimate of a plant's ability to absorb nutrients from the soil) and mycorrhizal formation suggests that a continuous range of mycorrhizal strategies apparently exists-ranging from species with coarse, long-lived roots of which almost all form mycorrhizas at one end of the continuum. to those with little or no mycorrhizas that have highly active and extensive roots at the other.

6. Mycorrhizas and Plant Responses to Pollution and Other Stresses Plants that are dependent on mycorrhizas require them to supply nutrients at adequate levels to sustain "normal" growth and reproduction where they occur in natural ecosystems (Section III.E.7), but mycorrhizas may have other les-direct influences on plant fitness and survival. The indirect mycorrhizal benefits that have been most often reported include increased tolerance to various biotic or abiotic stresses. Associations with ECM or VAM fungi have been reported to increase host resistance to pathogens (Chakravarty and Unestam, 1987; Dehne, 1986; Duchesne et al., 1987). This increased resistance may involve improvements to host plant mineral nutrition, physical protection of roots by mantle hyphae in the case of ECM, phytoalexin productionantimicrobial chemicals such as flavonoids or phenolics-or other mechanisms that are not well understood (Bagyaraj, 1984; Duchesne et al., 1987; Harley and Smith, 1983; Morandi et al., 1984). Davis and Menge (1980) were able to demonstrate disease-suppression benefits from a VAM association when mycorrhizal roots were isolated from diseased roots (in a split root experiment), which suggested that the beneficial effects of VAM were due to enhanced host plant mineral nutrition. Fungi forming ECM can be antagonistic to other microbes, perhaps by producing antibiotics (Garrido et a l . , 1982; Kope and Fortin. 1989; Rambelli, 1973). Kope and Fortin (1989) found that cell-free extracts from cultures of 7 out of 16 isolates of ECM fungi they screened inhibited the growth of many of the root-pathogenic fungi they tested and in some cases cell lysis was observed. Duchesne et al. (1989) reported that oxalic acid, which was produced by an ECM fungus (Paxillus involurus), inhibited a Fusarium pathogen of pine roots. There are some instances when mycorrhizal associations do not reduce disease severity. For example. Afek et al. (1990) observed a negative interaction with VAM (reduced root colonization) caused by several common soil pathogens when crop plants were grown in non-fumigated soil.

226

kl. I + K I I N I > K E l T

Interactions between nematodes and VAM are complex (see review by lngham. 1988). Plant-feeding nematodes generally reduce the growth of mycorrhizal plants, but they still usually grow better than non-mycorrhizal control plants. Plant-feeding nematodes generally avoid roots already containing VAM and vice versa (although in some cases the activity of VAM fungi and nematodes are mutually enhanced). Fungalfeeding nematodes can reduce root colonization or mineral nutrient uptake by the soil hyphae VAM fungi, but these effects may not be large enough to influence host plants. Nematodes and other mycophilous soil organisms can also have detrimental influences on ECM associations (Section III.B.3). There is a report that the fitness of insect larvae feeding on mycorrhizal plants can be reduced relative to those feeding on non-mycorrhizal plants (Rabin and Pacovsky. 1985). There apparently is little interaction between herbivore grazing of leaves and mycorrhizal colonization of roots (Allen e f al., 1989b; Wallace, 1987), although severe grazing can substantially reduce mycorrhizal colonization, presumably due to a reduction in photosynthate available to maintain root processes (Bethlenfalvay et a/., 1985; Borowicz and Fitter, 1990). Lapeyrie and Chilvers (1985) reported that a Eucalyptus species could grow in calcareous soil when associated with ECM but not with VAM. Mycorrhizal associations with certain isolates of VAM and ECM fungi can reduce hosts plant susceptibility to toxic metal ions (Denny and Wilkins, 1987; Dixon, 1988; Dueck et al., 1986; Jones and Huchinson, 1988; Koslowsky and Boerner, 1989; Schuepp et a/., 1987). However, mycorrhizal fungi themselves exhibit varying degrees of susceptibility to these factors and may enhance metal ion uptake in some cases (Smith, 1990). Ericoid mycorrhizal associations effectively “detoxify” metal ions and phenolic compounds which can occur in phytotoxic levels the acidic soils in which their host plants grow (Leake et al., 1989; Read 1983). Pesticides generally have little effect on mycorrhizal associations if used at recommended rates, but there are exceptions (Trappe et al., 1984). There is evidence that VAM associations can increase the uptake of soil pesticide residues by plants (Nelson and Khan, 1990). Acidic precipitation and atmospheric ozone and sulphur dioxide pollution have been implicated in the serious forest decline problems occurring in Europe and Eastern North America (Klein and Perkins, 1988; Smith, 1990). Forest decline is a complex process involving many biotic and abiotic factors, of which the disruption of nutrient cycling appears to be one of the most important (Klein and Perkins, 1988). In experiments, pollution in the form of O,, SOz,S dust, or acid rain can reduce VAM or ECM formation, but mycorrhizal plants were more tolerant to these stresses in some cases (Danielson and Visser, 1989;

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

227

Garret et al., 1982; Ho and Trappe, 1984; Reich et al., 1985; Shafer et al., 1985). Reductions in the biomass of feeder roots and mycorrhizas have been observed in forests where trees are declining (Mejstrik, 1989; Vincent, 1989), but it has not been determined if this is a cause or a symptom of tree decline. VAM colonization of roots may be associated with increased salinity tolerance in some saltmarsh plants (Rozema et al., 1986), but many plants in these habitats are nonmycorrhizal (Appendix 1). Graham and Syvertsen (1989) found that associations with a VAM fungus did not influence the tolerance of citrus seedlings to salinity, but enhanced chloride uptake. Pfeiffer and Bloss (1988) found that a VAM association increased plant growth in a saline soil, but so did the application of additional phosphorus. There is experimental evidence that mycorrhizal fungus hyphae can transport water, perhaps in sufficient quantities to help sustain plants during periods of water stress (Read and Boyd, 1986). However, the increased growth of mycorrhizal plants in dry soils is normally considered to occur by less direct mechanisms and often involves enhanced transpiration (Allen and Allen, 1986; Nelsen, 1987; Parke et a / . , 1983; Sieverding and Toro, 1988). Mycorrhizal increases to drought tolerance can be explained by improved phosphorus nutrition in many cases (Fitter, 1988; Nelsen, 1987), but there may still be transpirational differences between mycorrhizal and non-mycorrhizal plants of similar size and phosphorus content (AugC, 1989). In many of the experiments described above, mycorrhizal benefits to plants were assessed by contrasting the response of mycorrhizal and non-mycorrhizal plants to stresses at the same fertility levels, so it is often difficult to separate mycorrhizal benefits due to mineral uptake from other less-direct mechanisms (see Fig. 5 ) . While there may be more subtle indirect effects of mycorrhizas on plants, it seems that most of the increased stress tolerance of mycorrhizal plants can be directly related to enhanced uptake of mineral nutrients, especially phosphorus (Graham, 1987; Marschner, 1986; Nelsen, 1987).

7. The Value of Mycorrhizas to Plants in Natural Ecosystems The numerous factors which interact to determine if mycorrhizal associations are beneficial to plants in natural ecosystems or experiments are summarized in Fig. 5 . As was considered in Section III.E.5, the cost-effectiveness of mycorrhizal vs non-mycorrhizal root systems will largely depend on the availability of nutrients in soils and the relative expense of producing and maintaining root systems which are well suited for direct vs mycorrhizally mediated uptake (Table 5). Baylis (1975) was

j

228 A

M . BRLINDKE'M

*

F

Enhanced capture of soil nutrients with low mobility UDtake of water and mobile nutrients

*

Enerev costs due to mvcorrhizas

e

L

=I

+

Enerev costs due to extensivehctive roots

*

Improved resistance to stresses and disease ?

* *

Severe climatic or edaphic condition Mvcorrhizal inoculum limited Hvphal grazing or antagonism bv other ornanisms

B Soil nutrient levels (Soil factors influcncing nutricnt availability)

High

4

Abilitv of the fungus to S U D ~ I V nutrients (Inoculum levels, cflicacy of isolate used etc.)

Low

*Low

Ability of roots to absorb nutrients directly (Root system surl'acc area. plasticity. ctc.)

Low

*Low

High

4

High

lnternal dant nutrient reuuirements

(Adaptation to high or low nutrient supply)

Low

+Antagonisls Mvcorrhizosphere organisms absent Beneficial organisms absent absent

* b

*

*

I N1

._ c

15

Impact of absence of mvcorrhizae on control plants (Mycorrhizal depcndcncy of spccies)

Measured response to mvcorrhizal inoculation (Growth rcspnsc, incrcascd strcss rcsistancc, ctc.)

Fig. 5. Diagrammatic summary of factors with the potential to influence the effectiveness of mycorrhizal associations in natural ecosystems (part A: upper) and experimental systems (part B: lower). These factors are discussed in various sections of the review, but their relative importance is unknown. The last two factors in part B (lower) indicate that measured plant responses to mycorrhizas are relative to control plants grown without mycorrhizas (see text).

one of the first systematically to consider that these factors regulated variations between plants in their responsiveness to mycorrhizal colonization and the critical levels of nutrients, especially phosphorus, above which they no longer benefited from mycorrhizas. Evidence that there

MYC'ORRHIZAS IN N A T U R A L ECOSYSTEMS

229

are variations in the mycorrhizal dependency of plants has been provided by experiments where plant mycorrhizal responses are determined at soil nutrient levels comparable to those found in their natural habitats (Table 6). Many of the plants in Table 6 were found to be highly dependent on mycorrhizas. but these associations provided little o r no benefit in other cases. Despite having relatively extensive fibrous root systems, most sand dune and prairie grasses benefit from VAM associations, but grasses from other habitats may not (Table 6, Miller, 1987). Reports of non-mycorrhizal trees are rare and there is a substantial body of physiological evidence that trees normally are highly dependent on VAM or ECM associations (Fitter and Hay, 1987: Harley and Smith, 1983; Kormanik, 1981; Le Tacon et al., 1987; Lyr and Hoffman, 1967; Meyer, 1973). This may result because trees generally have relatively low root:shoot ratios and they often have short, heterorhizic lateral roots (Kubikova, 1967; Lyr and Hoffman, 1967). Salix nigra, a tree which grows in wet soils, is an exception to this rule, since it has elongated lateral roots, of which only a fraction develop ECM (Brundrett et al., 1990). Variations in mycorrhizal colonization resulting from changes in the nutrient status of plants/soils have been demonstrated in many experiments (Section III.E.4), but correlations between mycorrhizal colonization and plant or soil phosphorus contents are usually weak or non-existent in field trials (Abbott and Robson, 1991a; McGonigle, 1988) or natural ecosystems (Franklin and Harrison, 1985; Lee and Lim, 1989; McAfee and Fortin, 1989; Noordwijk and Hairiah, 1986). Boerner (1986) observed that more VAM colonization occurred in Geranium and Polygonarum spp. roots from less-fertile North American deciduous forest soils. Species in these genera have relatively finer root systems and lower levels of VAM when compared with many other plants in deciduous forests (Brundrett and Kendrick, 1988), so may have more facultative (and thus more variable) VAM associations. Correlations between soil nutrient levels and mycorrhizal formation in natural ecosystems would be expected to be uncommon, because natural soil nutrient levels are normally much lower than those reported to reduce mycorrhization in experiments, many plants are highly dependent over a range of soil fertility levels and plants and mycorrhizal fungi are adapted to the fertility of the soils in which they occur. It is an interesting paradox that plants in natural ecosystems are often less efficient at absorbing nutrients from soil than more opportunistic ruderal species or crop plants, but they generally also have lower requirements for nutrients such as phosphorus (Begg, 1963; Cardus, 1980; Chapin, 1988; Chapin et al., 1986). Herbaceous plants which grow

230

M. H R L I N D R E T T

Table 6 Mycorrhizal growth responses of plants from natural ecosystems

Ecosystem type

Mycorrhizal dependency

References

23+ ,5-

Janos (1980a)

+ (ECM) +

Berliner et nl. (1986) Jasper rt d.(1988, 1989~)

Plant ~

Tropical forest trees, herbs (28 spp.) Mediterranean shrub (Cistiis sp.) shrub (Acacia spp.) Broadleaved evergreen forest trees, herbs

+

Deciduous forest trees

+

shrubs ( 3 spp.) herbs (4 spp.) herb (Geruriiiirn sp.) Prairie grass (Schiznchyriutn sp.) grass (Andropogon sp.) warm-season grasses and forbs cool-season grasses forbs (3 spp.) Temperate arid ecosystems grasses (Agropyr-on spp.) shrub (Atriplex sp.) Alpine grasses Coastal sand dunes Grass (Urriola sp.) Disturbed habitats weed (Atnhrosia sp.) wild oats (Avetia sp.) Rushes and sedges (4 spp.) Aquatic plants submerged Rariirriculus sp.

+

+ -

+ +

Baylis (lY67), Hall (1977), Johnson (1977) Kormanik (1981). Kormanik et a / . (1982), Pope et a / . (1982). Tobiessen and Werner (1980) Sylvia (1987) Mosse (1978) Boerner (1990) Anderson and Liberta (1989)? Gerschefske Kitt et al. (1988) Gerschefske Kitt et ul. (1989), Daniels Hetrick rt a / . (1989) Daniels Hetrick and Bloom (1988)

-

Daniels Hetrick et a/. (1989) Anderson and Liberta (1987), Zajicek et d.(1986a)

-

+

Allen and Allen (1986). Loree and Williams (1987), Henkel et a/. (1989) Williams et a/. (1974)

-

Allen et a/. (1987)

+

+ +

Sylvia (1989)

-

Crowell and Boerner (1988) Koide et a/. (1988) Powell (1975)

-

Tanner and Clayton (1985)

Notes: Responses are to VAM unless otherwise stated; + substantial growth response to mycorrhizal inculation when grown at realistic nutrient levels; - little or no growth response to inoculation; k response varied with soil used.

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23 1

in fertile soils often have greater relative growth rates than plants from infertile sites (Grime, 1979; Tilman, 1988). Plants in natural ecosystems may be more tolerant of low nutrient levels because their slow growth results in less demand for nutrients. These plants also conserve resources within long-lived shoot and root structures, efficiently reclaim minerals from senescing tissues (Boerner, 1986; Chapin, 1988) and establish carefully regulated mycorrhizal associations (Brundrett and Kendrick, 1990b). Long-lived roots (with structural defensive features, such as an exodermis-(Section 1II.C.1) would initially be more expensive to produce, but thereafter confer increased resistance to adverse physical or biological soil factors. Despite their extremely fine lateral roots, plants with ericoid mycorrhizas are considered to be highly dependent on these associations, which are required to detoxify the soils in which they often occur (Leake et al., 1989). The achlorophyllous, subterranean protocorms of orchid seedlings are completely dependent on mycorrhizas but photosynthetic adult plants may be less so (Harley and Smith, 1983). Some adult plants with monotropoid, arbutoid, or orchid mycorrhizal associations lack chlorophyll and are completely dependent on these associations for their sustenance (see Section III.F.l). Most of our knowledge about mycorrhizal benefits comes from studies of crop and forage plants. Pasture plants often occur naturally in similar habitats, but many crop plants have been changed considerably by selection and breeding programmes. Wild oat (Avena fatua) is less dependent on mycorrhizas than cultivated oat (Avena sativa) because the former has a higher root/shoot ratio and adaptations to low nutrient levels such as inherently slower growth (Koide ef al., 1988). Other comparisons between domesticated plants and their wild relatives would be instructive, since selection of plants that perform well in agricultural conditions may well also have resulted in the opposite situationreduced mycorrhizal dependency. Many crop plants might be expected to be facultatively mycorrhizal because they have ruderal ancestors and were selected for rapid growth in high fertility soils. Mycorrhizal associations have been found to provide little benefit or even reduce the yield of some temperate field crops or pasture species (Jones and Hendrix, 1987; McGonigle and Fitter, 1988a), but many other cultivated plants receive a substantial benefit from these fungi (Hayman, 1987; Mosse and Hayman, 1980). Tropical crops and forage species often grown in acidic, highly infertile soils and many are highly dependent on mycorrhizas (Crush, 1974; Saif, 1987; Howeler et al., 1987). Morphological criteria (the proportion of a plants absorbing root system containing mycorrhizas) have been used to designate plants with different degrees of mycorrhizal dependency (Section III.D.2), but

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physiological evidence (the relative benefit provided by mycorrhizas at realistic nutrient levels) would provide a more consistent basis for these designations. Unfortunately, it is difficult to use physiological definitions during mycorrhizal surveys in natural ecosystems, because it usually is not possible to manipulate (and may be difficult to measure) soil nutrient availability. Plants in natural ecosystems belong to a continuum ranging from plants which consistently have mycorrhizas in almost all of their roots to those that never do (Fig. 4). Morphological and physiological definitions of mycorrhizal dependency would probably be in close agreement when applied to plants from either end of this spectrum. However, there is likely to be less agreement about the value of mycorrhizas to plants with intermediate levels of mycorrhizas, since there is little information about the proportion of a plant’s root system which is required to form an effective association (and this proportion may be influenced by soil and other environmental conditions). The use of fungicides which fairly selectively inhibit mycorrhizal fungi, provides another approach to analysing the benefit of mycorrhizal associations in natural communities. Benomyl application to soils can substantially reduce VAM formation, phosphorus inflow and shoot biomass in experiments (Borowicz and Fitter, 1990; Fitter and Nichols, 1988; Daniels Hetrick er al., 1989: Trappe et a l . , 1984). When applied to an alpine grassland community, benomyl reduced mycorrhizal infection, but had no consistent influence on plant phosphorus content (Fitter, 1986b). Gange er al. (1990) reduced mycorrhizal colonization and plant productivity by applying the fungicide Rovral (iprodione) to an early successional community. Koide ci al. (1988) found that fungicide application had little effect on VAM, but increased plant abundance in an annual plant community. Ideally, it would be possible to find a mycorrhizal inhibitor that could be applied to shoots and would be phloem-translocated to active roots, which would then remain free of mycorrhizas (Fitter, 1989). A fungicide of this type would have minimal influence on soil nutrient cycling processes or the mycorrhizal associations of untreated plants. Poor correlations between experimentally induced changes in mycorrhizal colonization levels and plant growth responses or phosphorus contents in field experiments may occur because: ( i ) mycorrhizal associations provided little benefit to the species involved or are only required during certain stage in their life history (Fitter, 1989), (ii) plants normally form more mycorrhizas than they actually need so that changes have little effect below a certain threshold (McGonigle, 1988). (iii) roots were not very active at the time of application, or (iv) the effects require long periods to develop because plants grow very slowly. A detailed cost-benefit analysis of a mycorrhizal association must

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balance productivity gains (resulting from enhanced mineral nutrient uptake) against the costs of maintaining these associations (Koide and Elliott, 1989). Root production and turnover rates are difficult to measure accurately but are thought to account for 40-85% of the net primary productivity of plants in ecosystems (Fogel, 1985; Hayes and Seastedt, 1986; Vogt et al., 1987). This energy is used to support root growth, respiration, exudation, as well as mycorrhizas and other associations (Lambers, 1987). By comparing mycorrhizal and phosphorusfertilized plants with similar relative growth rates, Baas et al. (1989b) found that mycorrhizal Plantmgo major sp. plants had 30% higher root respiration rates. Koch and Johnson (1984) found that 3-5% of “Clabelled photosynthate was supplied to mycorrhizal fungi in half of the root system in a split root experiment. Jakobsen and Rosendahl (1990) found that approximately 20% of the photoassimilate of cucumber plants with VAM was required for mycorrhizal events (using a “CO2 labelling experiment). It is considered that trees in boreal forests expend about 50% of their primary productivity below ground (Fogel, 1980, 1985; Vogt et al., 1982) and it has been estimated that 20-30% of this energy is supplied to ECM fungi to form mycorrhizas, hyphae, sclerotia and reproductive structures (Harley, 1971; Odum and Biever, 1984). However, these estimates of mycorrhizal costs are rather preliminary (because of inaccuracies in root turnover rate measurements and yearly variations in ECM fungus reproduction). In general it seems that VAM associations utilize 7-17% of the energy translocated to roots (Harris and Paul, 1987: Lambers, 1987), but ECM associations may require 20-30% of this energy (Harley, 1971; Odum and Biever, 1984). Nutrient uptake without mycorrhizal mediation also requires respiratory energy and in the case of two Carex species (most sedges are non-mycorrhizal: Powell, 1975) this can amount to 25 to 35% of their total respiration (Van der Werf et al., 1988). Mycorrhizal roots would still expend some energy on direct ion uptake, but these expenditures would almost certainly be much less than those of non-mycorrhizal species, which generally have more extensive and active root systems and higher root/shoot ratios than mycorrhizal plants (Table 5 ) . There is also some evidence that plants growing in full sunlight can have excess photosythetic energy that overflows into an alternative (cyanide-resistant) respiratory pathway where it is lost (Lambers, 1985). Cost-benefit analysis calculations based on photosynthetic energy would be less relevant if plants in natural ecosystems are able to capture photosynthetic energy that is in excess of their needs for well regulated growth, respiration and storage. When plants that would obtain little or n o benefit from mycorrhizas in natural ecosystems are compared with plants that are dependent on mycorrhizas, the former would normally

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M . BRUNDRETT

have root systems that are more expensive to maintain. However, when the costs vs benefits of one plant’s mycorrhizal associations are being considered the results may well depend on soil fertility levels and other factors. The benefit provided by mycorrhizal associations to plants has been established by numerous glass-house and growth-chamber experiments, but less often under field conditions and rarely in ecosystems (Abbott and Robson, 1991a; Fitter, 1985; Nelson, 1987). There has also been a tendency to use experimental conditions which would maximize mycorrhizal growth responses, without considering complicating factors (as shown in Fig. 5) which would reduce mycorrhizal efficacy (Fitter, 1985). Despite these criticisms, there is little doubt that mycorrhizal associations have an important role in nutrient uptake in natural ecosystems. This assertion is supported by the following indirect evidence: (i) The majority of plants growing in natural ecosystems have mycorrhizal associations (Appendix 1). (ii) Experiments with plants from natural ecosystems have demonstrated benefits from mycorrhizas at appropriate nutrient levels (Table 6). In places where mycorrhizas are absent, their reintroduction may result in increased plant productivity (Sections III.B.4, III.F.4). Application of fungicides to natural communities has resulted in measurable reductions in plant production or diversity, at least in some cases. As was considered in the previous section, mycorrhizal associations may reduce the impact of various plant stresses, largely through enhanced plant nutrition. (iii) Mineral nutrients (especially phosphorus and nitrogen) provide most of the beneficial effects of mycorrhizal associations observed in experimental studies (Hayman, 1983: Harley and Smith, 1983) and are amongst the most important factors limiting plant production in ecosystems (Coleman er al., 1983; Chapin e[ al., 1986; Kramer, 1981; Fitter, 1986a). Mineral nutrient constraints on plant growth are considered to be important in tropical forests (Janos, 1987; Jordan, 1985), deserts (Noy-Meir, 1985), mediterranean regions of South Africa and Australia (Jeffrey, 1987; Lamont, 1982), temperate deciduous forests (Wood et al., 1984), prairies ( R i s e r , 1985), boreal coniferous forest (Larsen, 1980) and arctic vegetation communities (Chapin and Shaver, 1985; Tieszen, 1978). In these regions many plants have strategies to ensure nutrient uptake and conservation, including long-lived evergreen or xeromorphic (desiccation resistant) leaves which thought to conserve nutrients as well as water (Chabot and Hicks, 1982; Fitter, 1986a; Jeffrey, 1987). (iv) In experiments radioactive tracers have demonstrated the rapid transfer of nutrients to roots, that most probably occurs through a

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235

network of mycorrhizal fungus hyphae (Section III.E.2). (v) The extensiveness, responsiveness and activity of a plant’s root system will determine its ability to obtain relatively immobile soil resources and these root characteristics are often negatively correlated with mycorrhizal colonization or mycorrhizal dependency (Table 5 ) . Some mycorrhizal species in natural ecosystems have coarse, relatively inactive roots that would be hopelessly inefficient at direct nutrient absorption. Comparing the biomass of similar lengths of roots and mycorrhizal fungus hyphae, suggests that hyphae are a considerably less expensive way of exploring the soil (Harley, 1989). (vi) As was considered above, host plants spend a significant proportion of their energy budget to support mycorrhizal fungi. Ecological factors (such as low light levels) that should limit the plant’s ability to supply photosynthate, if anything favour mycorrhizal species over those that are non-mycorrhizal (Brundrett and Kendrick, 1988).

F. The Ecology of Mycorrhizal Plants Interactions between plants which are mediated by above-ground processes include competition for light and space and variations in direct responses to the physical environment (Grime, 1979; Tilman, 1988), but it is difficult (and perhaps irrelevant) to completely eliminate soil factors and mycorrhizal fungal influences from consideration. Many mycorrhizal ecology topics are poorly understood, because plant ecologists rarely consider mycorrhizas and mycorrhizal investigations usually are not conducted in ecosystems (Fitter, 1985, 1989; Harley and Harley, 1987; St John and Coleman, 1983). However, much relevant information about both mycorrhizal partners alone, or in association has been obtained from experimental o r agricultural situations and is of value in predicting the response of mycorrhizas in ecosystems. Papers on mycorrhizas in ecosystems (especially regarding regeneration after disturbance or forestry) are now appearing in the literature with increasing frequency.

I . Mycorrhizas and Nutrient Cycling in Natural Ecosystems Mycorrhizal fungi are generally considered to be incapable of utilizing complex substrates such as cellulose and lignin as energy sources and to depend on their host plant for nutritional support (Harley and Smith, 1983). The intimate association of mycorrhizal roots with leaf litter has resulted in the hypothesis that mycorrhizal fungi could be directly involved in leaf litter decomposition, but there is no good evidence that

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M . RKIJNDKETT

this occurs in ecosystems (Harley and Smith, 1983). Some ECM fungi have the enzymatic ability to degrade cellulose and pectin, but this activity is generally much less than that of saprobes (Dahm et al., 1987; Harley and Smith, 1983; Haselwandter et al., 1987). It has recently been discovered that some ECM fungi can utilize organic (protein) sources of nitrogen (Abuzinadah et al., 1986; Abuzinadah and Read, 1986) and phosphorus (Haussling and Marschner, 1989; Ho, 1989). In soils where ECM trees occur, 50% of the phosphorus can be in organic forms that could be broken down by rhizosphere phosphatases produced by ECM fungi (Haussling and Marschner, 1989). Soil minerals are weathered by microbial activity as well as physical processes (Coleman et al., 1983; Robert and Berthelin, 1986). Hyphae of the ECM fungi Hysterangium sp. and Paxillus involutus apparently weather clay minerals by producing calcium oxalate (Cromack et al., 1979; Lapeyrie, 1988). Oxalates produced by ECM fungi may chelate iron to release phosphorus from iron phosphates (Cromack et al., 1979; Lapeyrie, 1988) and siderophore production has also been proposed as a mechanism for enhanced phosphorus absorption by VAM fungi (Bolan et al., 1987; Jayachandran et al., 1989). Reducing activity can be detected in the rhizosphere of ECM roots and may be an important mechanism to enhance the absorption of oxides of nutrients such as manganese (Cairney and Ashford, 1989). Ericoid mycorrhizal fungi apparently normally utilize organic sources of nitrogen (amino acids) and phosphate (Bajwa and Read, 1986; Read, 1983; Straker and Mitchell, 1986) and can degrade lignin, at least in some cases (Haselwandter et al., 1987). These specialized nutritional abilities of ericoid mycorrhizal fungi are likely to be particularly valuable in heathland soils where organic forms of nutrients, such as free amino acids, may predominate (Abuarghub and Read, 1988; Leake et a l . , 1989). The influence of mycorrhizal fungi on forms of soil minerals that are generally considered not to be available to plants requires further investigation. Monotropoid and some arbutoid and orchid mycorrhizal hosts have a highly reduced root system and are without chlorophyll, so must rely on mycorrhizal fungi to supply all of their needs (Furman and Trappe, 1971)-a reverse of most relationships between higher plants and fungi. Isotope tracer studies have demonstrated nutrient transfer to plants with monotropoid mycorrhizas from nearby trees, presumably through a common mycorrhizal fungus (Bjorkman, 1960; Newman, 1988; Vreeland et a l . , 1981). Plants with arbutoid mycorrhizas usually have chlorophyll, but also associate with the same fungi as adjacent trees (Molina and Trappe, 1982a). Members of the Gentianaceae that have VAM associations may not support further spread of the fungus (Jacquelinet-Jean-

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237

mougin and Gianinazzi-Pearson, 1983; McGee, 1985). Warcup (1988) found that Australian members of the genus Lobeliu (and its allies) were often highly dependent on VAM or ECM associations during seedling establishment and apparently required the presence of a companion plant for subsequent growth. These experiments with gentians and lobelias suggest that some plants in natural ecosystems greatly benefit from mycorrhizal associations, but provide little to the fungus in return, so these fungi (and to some extent the lobelia or gentian) would be supported by other members of their community. Orchid mycorrhizal fungi can transfer “C to their host plant especially when they are young (Alexander and Hadley, 1985), but the source of this carbon has not been determined. Endophytes isolated from orchid roots are often designated as Rhizoctonius-a diverse group of nonsporulating fungi which include plant pathogens as well as saprophytes (Hadley, 1982; Warcup, 1981; 1985) and similar fungi have been found to occupy hyphae and spores of VAM fungi (Williams, 1985). However, detailed studies of orchid endophytes suggest that many of these fungi may have fairly specialized associations with orchids (Currah er ul., 1987; Ramsay er al., 1987; Warcup, 1981). The nature of interactions between orchid mycorrhizal fungi and other plants in their native habitats warrants investigation. Orchid endophytes may be saprophytes, but it is possible that many of these fungi have a detrimental influence on other plants or their mycorrhizal fungi and this is almost certain to be the case with achlorophyllous orchids that require photosynthetically acquired energy from some other source. Arrnillurieflu rnelleu, a fungus which can be parasitic on plants, forms mycorrhizal associations with some chlorophyll-free orchids (Campbell, 1962; Terashita, 1985). It is interesting to note that chlorophyll-free and green individuals of the normally photosynthetic orchid Epipactis helleborine apparently had a common mycorrhizal fungus, which was able to support the growth of fully heterotrophic individuals (Salmia, 1988, 1989). It has been suggested that epiphytic orchids sometimes have a detrimental effect of their host trees (Johansson, 1977). Plants such as orchids, gentians and lobelias are renowned for their showy flowers, which are often disproportionately large. It may be that growth of these plants occurs partially at the expense of other members of the community (by way of their mycorrhizal fungi). However, these fascinating plants are usually not abundant in ecosystems so would have little impact on their fully autotrophic neighbours. Energy and nutrients from plants are recycled through six major pathways in ecosystems; (i) grazing, (ii) seed consumption, (iii) feeding on nectar, (iv) loss of soluble exudates, (v) active extraction by parasitic and mutualistic organisms and (vi) decomposition of plant structures

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M . HKUNDRETT

(Odum and Biever, 1984). The last process (litter decomposition) is the most important, but energy and nutrient flow through the other cycles is usually also significant (Odum and Biever, 1984). Decomposition processes in the soil involve many organisms in a complex food web, which culminates in nutrient absorption by roots (Coleman, 1985). This food web consists of hierarchies of microbes, microbe-feeding and predatory soil fauna, which form a pyramid with the more primitive organisms that feed on organic substrates at its base (Price, 1988). Measurements of root litter production in natural ecosystems indicate that below-ground litter (dead roots and mycorrhizas) can amount to 2-5 times aboveground litter production (Fogel, 1988; Raich and Nadelhoffer, 1989). It has been suggested that the presence of mycorrhizal roots can reduce the rate of leaf litter decomposition, perhaps by absorbing nutrients required by saprobic organisms (Cuenca et af., 1983; Gadgil and Gadgil, 1975). However, mycorrhizal roots had a greater influence on moisture levels in litter and their exclusion did not influence decomposition rates in other cases (Harmer and Alexander, 1985; Staaf, 1988). It seems likely that factors other than mycorrhizal fungus activity are responsible for the slow nutrient cycling which occurs in ECM tree dominated forests (Section 111. F.5). In a microcosm experiment, Dighton et al. (1987) reported increased rates of decomposition of some organic substrates when mycorrhizal roots were present and this effect on decomposition was reduced by the presence of a saprobic fungus. Hyphae from one mycorrhizal fungus are often associated with several different host plants in ecosystems to form a common pool of nutrients, but it is not known if all associated plants have equal access to this pool (Newman, 1988). It has been hypothesized that mycorrhizal fungus hyphae that are already present in senescing host roots may scavenge mineral nutrients before they escape into the soil solution or are absorbed by saprobes, and then transfer them to other associated roots, thus partially bypassing the soil food web (Newman, 1988; Newman and Eason, 1989). Dying plants roots can loose 60% of their nitrogen and 70% of their phosphorus within 3 weeks and much of this ends up in neighbouring plants (Eason and Newman, 1990). The magnitude of this rapid dying-living root nutrient transfer was substantially greater if cohabitating plants were mycorrhizal, providing evidence that the transfer occurs through VAM fungus hyphae links (Newman and Eason, 1989). As was considered above, mycorrhizal fungi are generally considered to enhance plant uptake of inorganic nutrients, but organic nutrient sources may also be important in some ecosystems and much of the soil nutrient pool is likely to be contained within the living biomass of organisms belonging to a complex decomposition food web. This likely

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results in competition between mycorrhizal fungi and other soil organisms for nutrients, but the nature and importance of this competition is unknown (St John and Coleman, 1983). It is possible that mycorrhizal fungi co-operate with other soil organisms to obtain nutrients that are spatially or chronologically separated from roots. There are seasonal variations in the availability of nutrients in natural ecosystem soils (Gupta and Rorison, 1975; Versoglou and Fitter, 1984), which may not always coincide with periods of maximum mycorrhizal root activity. Other soil microbes and non-mycorrhizal plants may help to prevent nutrient loss from the system at these times. Mycorrhizal roots or hyphae are often spatially associated with soil organic material in natural ecosystem soils (Section III.E.2). Efficient conservation of minerals, especially phosphorus, has been observed in deciduous forests, eucalyptus forests and tropical rain forests and is thought to result from the combined activities of microbes and roots (Attiwill and Leeper, 1987; Jordan, 1985; Wood et al., 1984). This efficient conservation suggests that the release of nutrients by the activities of saprobic organisms is tightly coupled with uptake by roots, bacteria and fungus hyphae concentrated near the soil surface (Wood et af., 1984). It is probable that mycorrhizal fungi have an important role in the later stages of this process, but this role has not been directly investigated.

2. Mycorrhizas and Plant Nutrient Competition The supply of resources such as light, water and nutrients often limits the growth of plants in natural habitats, but to obtain a greater proportion of one resource, plants must allocate more of their growth to the structures responsible for obtaining it, which may reduce their ability to compete for other resources (Tilman, 1988). For example, plants must balance their expenditures on stem and leaf growth required to obtain light against root costs associated with water and nutrient uptake. Root competition for soil resources is often more important than shoot interactions (Wilson, 1988) and root biomass in a particular site has been found to be inversely correlated with soil nutrient levels in natural ecosystems (Lyr and Hoffmann, 1967; Gower, 1987; Vogt et al. 1987). Plant root systems usually intermingle with those of other plants, so would normally be competing for the same soil resources (Brundrett and Kendrick, 1988; Caldwell, 1987: Caldwell and Richards, 1986; Chilvers, 1972; Kummerow, 1983; Richards and Caldwell, 1987). The relative availability of different soil resources can influence the outcome of competition between species with different capacities to obtain these resources (Tilman. 1982). Genetically determined characteristics of root systems that would influence the ability of plants to

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M . RRUNDREIT

obtain immobile soil nutrients such as phosphorus include total root length, rooting depth, geometry and plasticity (Section III.E.5). Root length and rooting depth relative to leaf transpirational area are particularly important for water acquisition, but other factors such as root xylem characterisitics and the proportion of roots which can absorb water are also important (Crombie et af., 1988; Eissenstat and Caldwell, 1988; Fitter and Hay, 1987; Hamblin and Tennant, 1987; Richards, 1986; Richards and Caldwell, 1987). Differences in root system responsiveness to temporarily available water or nutrients also influence the outcome of interspecific competition (Campbell and Grime, 1989; Franco and Nobel, 1990; Jackson and Caldwell, 1989). Co-existing plants in natural communities may avoid competition for nutrients by having roots which are active at different times of the year (Brundrett and Kendrick, 1988; Fitter, 1 9 8 6 ~Daniels ; Hetrick et al., 1989; Veresoglou and Fitter, 1984). The outcome of competition between plants with different mycorrhizal strategies will depend on the nature of soil resource(s) that are limiting plant productivity. Resources such as water and some mineral nutrients can move rapidly through soil by diffusion and would be obtained directly by roots even if they were non-mycorrhizal, while poorly mobile nutrients such as phosphorus are more efficiently obtained by fungus hyphae if plants are mycorrhizal (Section III.E.5). Thus, the outcome of competition between mycorrhizal and non-mycorrhizal plants should favour the former species if phosphate is limiting, but the extensive roots of non-mycorrhizal of facultative species should be more effective if water is in short supply. These differences in the efficiency of resource acquisition may be less important than other factors in communities where plant growth is restricted by severe edaphic or climatic conditions and non-mycorrhizal species are often common (Section III.F.5). When growing together, plants with the same type of mycorrhizal association may be more equal competitors than plants without mycorrhizas or with different types of mycorrhizas (Newman, 1988). Mycorrhizas should partially alleviate differences in the competitive ability of species to obtain immobile nutrients such as phosphorus by increasing the functional similarity of roots that differ in form. This may explain why species with highly mycorrhizal, coarse roots as well as those with extensive, fine root systems and lower levels of mycorrhizas successfully co-exist in habitats such as deciduous forests and prairies (Brundrett and Kendrick, 1988; Daniels Hetrick and Bloom, 1988). However, in these communities plants belonging to these two contrasting groups also avoid competition by having periods of maximum root activity at different times of the year. Different forms of competition may occur between plants which differ

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

24 1

in their mycorrhizal dependency, but there is little available experimental evidence to support these suggestions. Fitter (1977) reported that the addition of VAM to two grasses increased dominance of Holcus lanatus over Lolium perenne, when compared to root competition alone. Crush (1974) and Hall (1978) found that VAM increased the growth of clover grown in competition with ryegrass (which suffered from nitrogen deficiency) at low soil phosphate levels. In experiments with Agropyron species, which receive little benefit from mycorrhizas, these associations had little impact on the outcome of competition with other species (Allen and Allen, 1986). Daniels Hetrick et ul. (1989) studied competition between two prairie grasses- Andropogon gerardii, a species which was highly dependent on mycorrhizas (in the soils used) and Koeleria pyramidatu, which did not respond to VAM. They found that the outcome of competition favoured A . gerardii when mycorrhizas were present or phosphorus was added, while K . pyramidata was only successful when mycorrhizas were suppressed and phosphorus levels were low. In natural ecosystems the diversity of host plants is often substantially greater than the diversity of mycorrhizal fungi and there is no reason to believe that individual plants could maintain separate networks of mycorrhizal hyphae. Thus it would be normal for different individuals, species, ages or growth forms of plants to be interconnected by one or more mycorrhizal fungus (Harley and Smith, 1983). It has been suggested that these mycorrhizal hyphae networks could reduce plant nutrient competition and help support subordinate species (such as seedlings, shaded by a canopy of mature plants) by interplant nutrient transfer (Francis et al., 1986; Grime et al., 1987; Newman, 1988; Ocampo, 1986). In the experiments described above, mycorrhizal influences on the outcome of competition do not seem to involve nutrient transfer (or ryegrass would have obtained nitrogen from clover in Crush’s (1974) or Hall’s (1978) experiments). Grime et a f . (1987) demonstrated increased diversity (survival of subordinate species) in a microcosm experiment when species were growing with VAM fungi. They proposed that interplant nutrient transfer (measured by “ C ) was an important mechanism for maintaining higher plant diversity in this system. However, Bergelson and Crawley (1988) and Newman (1988), consider it more likely that this microcosm experiment demonstrated that the subordinate species had a greater requirement for mycorrhizas so their growth was increased relative to that of dominant grass plants when mycorrhizas were present in the system, while the grass plants (which have fine root systems) were at a greater competitive advantage when mycorrhizas were withheld. In separate experiments using sorghum and Plantago, Ocampo (1986) and

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Eissenstat and Newman (1990) could detect little P-transfer between mature plants and seedlings of these species when they shared the same VAM fungus and competition with mature plants adversely influenced the seedlings. Mycorrhizas may reduce nutrient competition between plants by some interplant nutrient transfer (section III.E.2). but they also increase the functional similarity of roots and this later role would appear to be more likely to help maintain plant diversity in ecosystems. It seems likely that competition occurs between plants connected to a common network of mycorrhizal fungus hyphae for nutrients obtained by that fungus (Newman, 1988). It might be expected that the outcome of this type of competition would depend on the energy a particular host invests on mycorrhizal formation, since exchange across mycorrhizal interfaces occurs simultaneously in both directions. Thus plants which support more active mycorrhizal exchange sites in their roots (arbuscules, Hartig nets etc.) should be at a competitive advantage, but subtle differences in the efficiency or compatibility of different host-fungus combinations may also be a factor. The level of support plants provide to mycorrhizal fungi (in the form of root carbohydrates) is thought to be an important regulator of mycorrhizal formation (Section III.E.4.a). However, some heterotrophic or partially heterotrophic plants have unexplained mechanisms that allow them to obtain inorganic and organic nutrients from hyphal networks in sufficient quantities to sustain their growth, without providing any reciprocal benefit to the mycorrhizal fungus (Section III.F.1). A wide range of different ECM fungi may be present in a community and some of these fungi preferentially associated with a certain host (Section I1I.D. l ) , suggesting that interactions between mycorrhizal fungi may sometimes be a factor in plant nutrient competition. Perry et al. (1989) conducted competition experiments using two different coniferous host trees growing together with different combinations of FCM fungi. In these experiments mycorrhizal fungi greatly reduced competitive effects by increasing phosphorus uptake by both species. In another microcosm experiment. Finlay (1989) grew seedlings of Pinus sylvesrris and Larix eurolepis together with one of three ECM fungi, two of which are considered to be specific associates of Larix spp. Growth of Pinus was found to be substantially better when the non-specific ECM fungus was present than when Larix-associates were used, while these latter fungi resulted in the best growth and phosphorus uptake by Larix seedlings. This experiment provides evidence that the presence of host-specific fungi can shift the balance of nutrient competition in favour of their hosts. The outcome of nutrient competition between mycorrhizal and nonmycorrhizal plants has also rarely been considered. In a study of early

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successional species, Crowell and Boerner (1988) found that non-mycorrhizal Brassica plants could have a greater influence on mycorrhizal Ambrosia plants than other Ambrosia neighbours, but this effect was not consistent. The presence of VAM fungi can increase the growth of host plants (grasses) in the presence of non-mycorrhizal competitors (Allen and Allen, 1984; Ocampo, 1986). In studies of this type it can be difficult to separate the competition for resources by roots and mycorrhizas from the growth benefits provided by mycorrhizas (when compared to growth when mycorrhizas are withheld-see Fig. 5 ) . Other factors such as allelopathy (Section III.F.3.a), that way influence the outcome of competition between mycorrhizal and non-mycorrhizal species, also need to be considered. It has been suggested that mycorrhizal fungi may have an adverse effect on non-mycorrhizal plants that is not related to nutrient competition (Allen and Allen, 1984). The sequestering of nutrients in mycorrhizal fungus hyphal pools may provide a competitive advantage to highly mycorrhizal over those with other nutrient uptake strategies (Newman, 1988). Root systems must evolve in response to the environmental factors most limiting to plant growth (Caldwell, 1987; Fitter, 1986a; Tilman, 1988). However, these factors may be in conflict (for example a root system optimized for water uptake would be very different than one that is most efficient at mycorrhizal formation), so it is not surprising that results of evolutionary processes have often produced plants with substantially different root systems (strategies) even in the same habitat. These differences in root strategies may influence the outcome of competition between species and ultimately the composition of plant communities.

3. Other Mycorrhizally Mediated Interactions Between Plants In addition to their roles in nutrient cycling and competition for soil resources, mycorrhizal associations may be involved in other interactions between co-existing plants. The presence of plants which are strongly dependent on VAM can substantially increase mycorrhizal colonization of more facultative plants that would otherwise form little VAM (Hirrel et al., 1978; Miller et al., 1983; Stejskalova, 1989). However, it is known if this type of mycorrhizal enhancement is detrimental or beneficial to facultative species, or if this depends on soil fertility. Attempted colonization of non-host roots by mycorrhizal fungi can result in wounding responses which may have an adverse influence on these plants (Allen et al., 1989a). ( a ) Allelopathic interactions involving niycorrhizal fungi. Chemical substances liberated in soil by plants can have adverse effects on other

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plants (Rice, 1984) and these amensalistic interactions may involve mycorrhizal fungi (Perry and Choquette 1987). Leachates from the leaves of some plants (Cote and Thibault, 1988; Iyer, 1980; Rose et al., 1983) and lichens (Brown and Mikola, 1974; Fisher, 1979; Goldner ef al., 1986) can inhibit ECM fungi. However, Pteridium aquilinum (bracken), Heliunrhus occidentalis and Salsola kali, which are considered to be allelopathic. have little influence on the mycorrhizas of other species (Acsai and Largent, 1983; Anderson and Liberta, 1987; Schmidt and Reeves, 1989). Roots of Calluna vulgaris, a species with ericoid mycorrhizas, produce factors inhibitory to fungi forming other types of mycorrhizas (Robinson, 1972). The growth of plants with ericoid mycorrhizas can result in the accumulation in soils of phenolic compounds (as well as a low pH and metal ions) which can be toxic to plants and mycorrhizal fungi but are detoxified by ericoid mycorrhizal endophytes (Leake et al., 1989). Tobiessen and Werner (1980) and Kovacic et al. (1984) have suggested that chemical properties of leaf litter from Pinus trees with ECM may inhibit VAM fungi in soils under these trees. Chemical properties of leaf litter produced by trees may specifically inhibit some ECM fungi allowing other more tolerant fungi to become dominant (Perry and Choquette, 1987). There is some evidence that ECM fungi can produce substances that inhibit the growth of other ECM fungi (Kope and Fortin, 1989). The roots of non-mycorrhizal species apparently contain chemical factors inhibitory to mycorrhizal fungi (Section III.E.4.c), so these substances could adversely influence mycorrhizal formation in other species if released into the soil. The previous growth of non-mycorrhizal plants in a soil has been observed to have a detrimental effect on the subsequent infectivity of VAM fungi, in some cases (Baltruschat and Dehne, 1988; Hayman ef al., 1975; Iqbal and Qureshi, 1976; Morley and Mosse, 1976; Powell, 1982), but not in others (Ocampo and Hayman, 1981; Schmidt and Reeves, 1984; Testier et al., 1987). Ferulic acid, an allelopathic agent produced by Asparagus officinalis, has an inhibitory effect on VAM fungus activity, even though this plant normally benefits from mycorrhizas (Wacker er al., 1990). Roots contain many secondary metabolites that have the potential to have allelopathic influences on plants and their mycorrhizal associates, but the role of these substances in soils is very difficult to resolve (Section III.E.4.c). Allelopathic interactions between mycorrhizal and non-mycorrhizal species, plants with separate types of mycorrhizas, or trees with different populations of ECM fungi may be one of the factors that influences the composition or stability of plant communities.

4. Mycorrhizas and Plant Succession Successional changes to plant populations occur during ecosystems

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recovery from disturbance or establishment on new substrates (Barbour et al., 1987; Grime, 1979). During this process opportunistic (ruderal) species are gradually replaced by more specialized plants, as competition for space and soil resources becomes more important (Grime, 1979). Ruderal species, which are usually rare or absent from undisturbed sites (Grime, 1979), apparently often have root systems that would make them less dependent on mycorrhizas (Table 5 ) than climax vegetation species. The proportion of mycorrhizal roots has been observed to increase along with plant cover during succession in several natural ecosystems (Khan, 1974; Lesica and Antibus, 1985; Miller et al., 1983; Pendelton and Smith, 1983; Red and Haselwandter, 1981; Rose, 1988). In tropical forests (Janos, 1980b) and arid shrub/grass communities (Allen, 1984; Allen and Allen, 1984; Miller, 1979, 1987; Reeves et al., 1979) the first colonizers of disturbed sites are often non-mycorrhizal or facultative species, while obligately mycorrhizal plants became dominant later in succession. Miller (1987) has suggested that obligately mycorrhizal plants are more likely to be found where soil nutrient levels are low and disturbance is minimal, while moderate disturbance and soil nutrient levels favour facultatively mycorrhizal species and a combination of severe disturbance and high nutrient levels will favour non-mycorrhizal plants (severe disturbance and low nutrient levels tends to eliminate plants altogether) (Grime, 1979). There is evidence that succession in many communities follows these generalized trends, but there are also many exceptions. In arid regions non-mycorrhizal plants may dominate severely disturbed sites and the proportion of mycorrhizal plants may only reach previous levels after 20-30 years, but in other regions with more mesic climates mycorrhizal plants are often common during the initial stages of succession (Miller, 1987). In the temperate deciduous forest region plants with VAM such as Solidugo and Aster are important in early succession and non-mycorrhizal species only become dominant after massive fertilizer or biocide applications (Medve, 1984). Gange et al. (1990) found that the inhibition of VAM formation by the application of a fungicide significantly reduced the growth and recruitment of some species during the first year of succession after cultivation. In some arid habitats mycorrhizas may provide little benefit if the establishment of early successional species with low mycorrhizal dependency normally precedes successful establishment of mycorrhizal species (Allen and Allen, 1988; Loree and Williams, 1987). In coastal sand dunes most colonizing species form VAM and benefit greatly from this association, even though propagules of VAM fungi must often be in short supply (Koske, 1978b). Apparently propagules of both host plants and VAM fungi are codispersed to coastal habitats as plant colonization begins (Koske, 1988; Koske and Gemma, 1990). During succession in prairie

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ecosystems the proportion of mycorrhizal roots (but not total colonization levels) can decline during succession as fine roots (which are poorly colonized) become more prevalent (Cook et al., 1988). In deciduous forests, annuals and other ruderal (weed-like) species which are normally present only in disturbed sites (Rogers, 1982). tend to be non-mycorrhizal or have facultative VAM associations (Brundrett and Kendrick 1988). These relatively opportunistic species respond to openings in the tree canopy more rapidly than obligately m ycorrhizal species, which generally have slow growth rates and lower reproductive potentials, but mycorrhizal species gradually regain dominance. Janos (1980b) observed similar trends in tropical forests and found that early successional species often had lighter seeds (aiding long range dispersal) than obligately mycorrhizal plants from climax communities (where seedling survival would be more important). In boreal coniferous forests ECM inoculum levels decline rapidly after clear-cutting and changes to soil properties can inhibit surviving fungi, thus preventing tree regeneration (Section III.B.4). In these forests a succession of increasingly specialized fungi parallels re-establishment of host trees and soil conditions (Section III.B.6). It has been proposed that the absence of ECM inoculum may be one of the factors responsible for the absence of host trees in temperate grasslands (Barbour et af., 1987). While it is true that propagules of ECM fungi are gererally absent from prairie soils (Meyer. 1973; White, 1941; Wilde, 1954), other factors such as low water availability and frequent fires are considered to be responsible for the absence of trees (Risser, 1985). Abundant VAM fungus inoculum is normally present (Liberta and Anderson. 1986; Miller, 1987), but poor growth of trees with VAM associations has also been reported in prairie soils (White, 1941; Wilde, 1954). In some boreal coniferous forests the soil does not contain VAM inoculum (Kovacic et al., 1984; Tobiessen and Werner, 1980). Seedlings of some trees may initially have VAM associations when growing in disturbed sites, but generally only form ECM when they are mature (Gardener and Malajczuk, 1988; Lapeyrie and Chilvers, 1985). This mycorrhizal flexibility may help them colonize sites where inoculum of ECM fungi is absent, but suggests that mature plants are more specific in their mycorrhizal requirements than seedlings. In boreal forests species with VAM and nitrogen fixing shrubs with ECM are important during early succession before ECM trees regain dominance and may help maintain mycorrhizal fungi (Cromack, 1981; McAfee and Fortin. 1989). The impact of soil disturbance, which may precede succession, on niycorrhizal fungus propagules was considered in Section III.B.4. After the last ice age, the revegetation of glaciated regions i n North America and Europe was a gradual process (Davis, 1981; Delcourt and

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Delcourt, 1987). Initially tundra was replaced in North America by Picea spp. forests which dominated until Larix, Abies, Betulu and Pinus species migrated northwards and finally trees from present day deciduous forests (such as Fraxinw, UImus and Quercus, which were early migrants, and Acer, C'urya and Fagus, which were slower) extended to their present ranges. From our knowledge of the present day mycorrhizal associations of these trees, it seems that postglacial succession initially resulted in the dominance of trees with ECM, while trees with VAM were slower to become dominant or codominant. These trends may have resulted because, the prevailing climatic conditions favoured trees with ECM associations, VAM fungi were slower to disperse than ECM fungi, VAM host trees had more limited dispersal than trees with ECM, or a combination of these factors were involved. The early Pleistocene occurrence of fossil VAM fungus spores (Pirozynski and DalpC, 1989) suggests that inoculum dispersal was not the most important limiting factor. There is evidence that in some situations mycorrhizal inoculum levels may be one factor influencing the rate of succession, but changes to soil properties, or root system characteristics of plants may also be important and the gradual replacement of non-mycorrhizal species by those with VAM or ECM does not always occur during this process. While, successional processes often result in the increasing dominance of obligately mycorrhizal plants, there are many exceptions to this generalization and facultatively mycorrhizal or non-mycorrhizal plants may remain dominant or codominant in communities where severe edaphic or climatic conditions prevail, as will be considered in the next section.

5. The Distribution of Plants with Different Mycorrhizal Strategies A compilation of reports on the mycorrhizal status of selected plants in major world ecosystems and edaphically limited vegetation communities is presented in Appendix 1. Trappe (1987) has calculated that about 3% of angiosperms have been examined for mycorrhizal associations (coverage of the pteridophytes would be similar, but a much higher proportion of the gymnosperms have been examined). Mycorrhizas have been most studied in temperate ecosystems, especially in the northern hemisphere, where there have been numerous studies especially of dominant trees (there are more than 75 reports concerning Pinus syfvestris or Fagus syfvatica (Harley and Harley, 1987). but in other areas sampling has been sparse (in tropical forests very high plant diversity makes this almost impossible) (Janos, 1987). From the evidence presented in Appendix 1 and other compilations of the mycorrhizal literature (Harley and Harley, 1987; Kelly, 1950; Meyer, 1973; Newman and Reddell,

'

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1987; Trappe, 1987) correlations between environmental or soil conditions and the distribution of plants with different types of mycorrhizas can be established. Less is known about the occurrence of species with low or highly variable levels of mycorrhizas, since many separate reports are required to determine if a species is inconsistently mycorrhizal (Trappe, 1987). Plants that can form more than one type of functional mycorrhiza are rare in most ecosystems (Section III.D.2). The available evidence is more than sufficient to conclude that plants in most ecosystems are predominantly mycorrhizal and this fact is now as well established as many of the other assumptions on which science is based. Trappe (1987) has compiled information about the mycorrhizal status of angiosperms from the mycorrhizal literature to allow comparisons with plant growth forms and ecological functions (Fig. 6). In general there are few correlations between mycorrhizal strategies and plant life history (annuals vs perennials) or growth form (herbaceous plants, trees, etc.). Exceptions to this rule include the fact that trees and shrubs are much more likely to have ECM associations than annuals and herbaceous plants (Harley and Smith, 1983). Parasitic plants usually are

Fig. 6 . Data from a mycorrhizal database compiled by Trappe (1987) is charted to illustrate mycorrhizal trends amongst the Angiosperms as a whole and within some specialized groups of plants (numbers of species examined for each category = 6507, 674, 679, 455, 293. 53, 66. 84 respectively). The question mark for epiphytes indicates that sampling of this group has been very limited and biased. "Several" refers to species which have been reported to have more than one type of mycorrhizas, while other terminology is the same as that used in Fig. 2 .

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non-mycorrhizal (Currah and Van Dyk, 1986; Harley and Harley, 1987; Lesica and Antibus, 1986): some plant families, such as the Scrophulariaceae, contain both mycorrhizal non-parasitic and non-mycorrhizal parasitic members, the latter having probably evolved from the former (Alexander and Weber, 1984). The proportion and importance of plants with different mycorrhizal strategies in a North American hardwood forest community are shown in Fig. 3. In this community, the mycorrhizal relations of plants is more closely related to root phenology than plant growth forms or above-ground phenology. Epiphytes, plants which grow attached to tree branches or rocks rather than in soil, are most common in humid tropical sites (Benzing 1973). Limited surveys of epiphytic Orchidaceae have found many to have sporadic infection by orchid mycorrhizas (Benzing, 1982; Hadley and Williamson, 1972). The mycorrhizal status of other important groups of epiphytes-Filicales, Araceae, Bromeliaceae, Gesneriaceae etc. has largely remained unexplored, although there is a report of ECM roots within bromeliad leaf-base tanks (Pittendrigh, 1984). A shortage of mycorrhizal inoculum can occur during early succession (Section III.F.4), or in habitats where host plants normally do not occur (Berliner ef a / . , 1986; White, 1941; Wilde, 1954), but it seems that factors influencing the distribution of plants with a particular type of mycorrhizal strategy are usually more important. However, the distribution of individual mycorrhizal fungi could influence the occurrence of plants which have relatively specific associations with these fungi, as is the case with some terrestrial orchids (Ramsay ef al., 1987; Warcup, 1981). Vegetation on the Earth's surface can be classified into a number of distinct ecosystems with characteristic climates, soils and vegetation (Walter, 1979). These natural ecosystems occupy different parts of the Earth's land surface as a result o f variations in the climatic factors with the greatest influence on plants- temperature and moisture availability (Billings, 1974; Osmond et uf., 1987; Whittaker, 1975). Thus tropical vegetation changes along a moisture gradient extending from rain forests to deserts, while temperate ecosystems follow a similar gradient from deciduous forests to grasslands and cool deserts (see Fig. 7). Other factors which can influence the distribution of vegetation include competition (which is most extreme in moist and warm regions and least severe in extremely cold or dry habitats) excessive radiation in arctic or alpine sites, localized edaphic factors and human activities such as agriculture and forestry (Barbour er a / . , 1987; Osmond ef al., 1987). One ecosystem type may occur in widely separate regions or continents with similar climates (Walter, 1979). Despite overall similarities in growth form, major taxonomic differences often occur between the

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I

WET

SEASONALLY DRY

DRY

I

Fig. 7. Low temperature and drought stress arc thc two most important factors responsihlc for tlic distribution of world ecosystem types. Competition and other biological interactions are also important a n d would have the grcatcst intlucnce i n warn1 and/or wet ecosystems and less important in cold o r dry hahitats. This figure is hascd on similar charts in Billings (1974). Osrnond L’I NI. (1987) and Whittaker (1975).

vegetation in these aeparate floristic regions (Takhtajan, 1986). Because of these differences in host genetic background. generalizations about the inycorrhizal status of plants in one floristic region should not be indiscriminately applied to another. However, the relative importance of host genetic background and environmental factors i n determining

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25 1

mycorrhizal strategies may be elucidated by their comparison. In general, it appears that similar mycorrhizal strategies occur throughout an ecosystem, even though they often include plants which are widely separated taxonomically or geographically (Appendix 1). Thus environmental influences would appear to be more important determinants of mycorrhizal strategies than host taxonomic relationships or geographic barriers to host and fungus dispersal. Within an ecosystem, variations in soil conditions, such as excess water and salinity, can cause localized changes in community structure (Etherington, 1982; Whittaker, 1975). Mycorrhizal strategies in edaphic communities have been separately considered in Appendix 1, in an attempt to gain some understanding of the influence of gradients in soil factors or climatic conditions on these associations. Care must be taken in drawing conclusions from correlations between these gradients and the mycorrhizal relations of plants in natural ecosystems, because of the limits to our current knowledge of the environmental physiology of mycorrhizas. Excess water and/or poor drainage in wetlands and aridity or excessive salinity in arid regions can restrict root growth (Gregory, 1987) as well as influencing the composition of plant communities (Etherington, 1982). Vegetation in temperate, arid, tropical or arctic regions has been observed to have less mycorrhizas where soils are wet (Anderson et al., 1984; Khan, 1974; Miller and Laursen, 1978; Read et al., 1976 etc.), but plants with VAM or ECM are usually present and some plants that are totally emersed in water can still form VAM (Appendix 1). The mycorrhizal inoculum potential of soils can be much reduced by flooding, as is the case with VAM after rice culture (Ilag et al., 1987; Nopamornbodi et al., 1987). Mejstrik (1965) observed that most VAM activity in the roots of plants in a wetland community occurred at the times when the water table was lowest. The occurrence of plants with ericoid mycorrhizal associations is thought to be restricted in waterlogged soils, where plants with air-channels in their roots (such as non-mycorrhizal members of the Cyperaceae) become dominant (Leake et a l . , 1989). Kim and Weber (1985) found that mycorrhizas declined in sites with higher salinity and were absent in the centre of salt playas where salinity levels were extremely high. Members of the plant family Chenopodiaceae frequent saline soils (Etherington, 1982) and are usually nonmycorrhizal (Harley and Harley, 1987; Trappe, 1981), although some woody members of the genus Atriplex have VAM (Williams et al., 1974). Soil parent material composition influences soil chemical properties and vegetation types (Jeffrey, 1987; Proctor and Woodell, 1975). Soil

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pH and calcium levels can also influence plant distribution since some species (calcicoles) prefer alkaline, calcareous soils while others (calcifuges) characteristically occur on acidic soils where metal ions may occur at toxic levels (Kinzel, 1983; Rorison and Robinson, 1984). In alpine vegetation communities plants growing on crystalline substrates had significantly lower levels of mycorrhizas than plants growing on calcareous substrates (Lesica and Antibus, 1985). In alpine or arctic regions many plants growing in rocky soils were non-mycorrhizal, while plants growing in soils with more organic matter were more likely to have VAM (Currah and Van Dyk, 1986; Miller, 1982b). Soils based on serpentine rock parent materials often have distinctive vegetation and may have toxic metal ion levels or lower nutrient levels (Proctor and Woodell, 1975). Plants in a serpentine grassland in California were mostly well colonized by VAM (Hopkins, 1986). Further investigations are likely to unearth other correlations between soil properties and plant mycorrhizal strategies, but these factors apparently have a much greater direct influences on plants. There are many reports of mycorrhizal plants in arid ecosystems (Appendix l ) , but within these habitats gradients of decreasing soil moisture availability can be correlated with reductions in the proportion of mycorrhizal plants (Selavinov and Elusenova. 1974; Schmidt and Scow, 1986). Mycorrhizal activity in arid regions may also occur at greater soil depths than is usual in other habitats (Virginia et al., 1986; Zajicek et a l . , 1986b). There are substantial reductions in the proportions of plants with any kind of mycorrhizal association in high altitude or high latitude sites where cold climatic conditions prevail (Christie and Nicolson, 1983; Dominik et a l . , 1965a; Haselwandter, 1979; Read and Haselwandter, 1981; Trappe, 1988). Christie and Nicolson (1983) observed that some plants had VAM on sub-Antarctic islands but no mycorrhizal plants were found o n continental Antarctic sites and they suggest that inoculum of VAM fungi may be lacking from the latter sites. Plants with VAM can be uncommon relative to those with ECM or ericoid associations, or non mycorrhizal roots in high arctic habitats (Appendix l ) , perhaps because the short growing season, or very slow nutrient cycling limit the effectiveness of VAM in these habitats (Kohn and Stasovski, 1990). The root length (relative to leaf surface area) of alpine vegetation is greater at higher altitudes and may compensate for lower mycorrhizal levels (Korner and Renhardt, 1987). It would seem that the climatic gradients that have resulted in the formation of ecosystems (Fig. 7) have had little influence on the distribution of mycorrhizal associations as a whole (especially VAM which occur in almost all habitats), except that increasing severity of soil

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or environmental conditions is often correlated with a gradual decline in the importance or mycorrhizal plants relative to those that are non-mycorrhizal. There is evidence that mycorrhizal roots are more efficient at nutrient capture than non-mycorrhizal roots in most communities (Section III.E.5). In communities where plant productivity is severely restricted by environmental factors, photosynthetic energy capture is less likely to be limited by above-ground competition for light, while soil resources are often in short supply. Apparently this excess in photosynthate production relative to nutrient supply allows plants with diverse capture strategies-including extensive non-mycorrhizal or proteoid roots, carnivory, or parasitism, to compete more effectively with mycorrhizal plants. Indeed, plants with these specialized nutrient strategies are more common in communities such as mediterranean shrublands, where nutrient supply is severely limited (Jeffrey, 1987; Lamont, 1982). More specialized types of mycorrhizas (Ericoid, Orchid, Monotropoid etc.) are present in most ecosystems but are rarely dominant (Appendix 1). Ericaceous shrubs (presumably with ericoid mycorrhizas) are common in the understorey of many dry, nutrient-poor forests in North America (Hicks and Chabot, 1985). Plants with ericoid mycorrhizas are more likely to be abundant or dominant in soils which are acidic and very nutrient deficient (Leake et a l . , 1989; Read, 1983). Ericoid mycorrhizal fungi can greatly increase the tolerance of host roots to the toxic phenolic compounds and metal ions that occur on these soils (Leake et a l . , 1989). Non-mycorrhizal plants or those with other types of mycorrhizas would be at a disadvantage in these communities if they cannot tolerate these soil conditions, or utilize organic nutrient sources (Section III.F.l). Litter from plants with ECM o r ericoid mycorrhizal associations, apparently influences soil properties in ways that would be deleterious to plants with other other root strategies (Section III.F.3.a), which suggests that these plants may form communities which have a tendency to be self-perpetuating. Trees are the dominant form of vegetation in habitats where they are not excluded by adverse environmental conditions (aridity or low temperatures) or disturbance. Forests containing mixtures of tress with VAM and ECM associations are generally less common than those where trees with one type of mycorrhizas predominate and non-mycorrhizal trees have rarely been observed (Appendix 1). Trees with VAM occur in many families, while those with ECM belong to a smaller number of diverse families (Harley and Smith, 1983; Malloch et a l . , 1980). Lodge (1989) compared VAM and ECM formation in Populus and Salix (trees which can form both types of associations) over a range of natural and experimental soil moisture levels. He found that VAM

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associations were more common in field soils that were drier or wetter than those in which ECM predominated, although VAM activity was also found to be optimal in soils that were not flooded or dry in experiments. These results provide evidence of antagonistic interactions between mycorrhizal associations since VAM was only dominant where ECM activity was low (Lodge, 1989). Trees with ECM are dominant in forests with relatively low plant diversity at high longitudes or altitudes where cool temperatures prevail (Appendix 1; Harley and Smith, 1983; Read, 1983; Singer and Morello, 1960). It has been suggested that only trees with ECM can grow in these forests (Tranquillini, 1979). However, this restriction is not due to the mycorrhizal fungus, since herbaceous plants with VAM are present in these forests and well beyond the tree line in arctic and alpine communities (Appendix 1). Similarly, the lack of trees above the tree line does not result from the absence of ECM fungi which still form associations with shrubs in much more severe sites. Trees with ECM dominate most deciduous forests in Europe (Appendix 1). In North America, trees with ECM are more likely to be dominant in sites in warmer and dryer regions or where soils are poor (ridgetops, sandy glacial till) (Chabot and Hicks, 1982). In Europe, management practices have resulted in the predominance of Fagus sylvatica (Walter and Breckle, 1985), an ECM tree, while in northeastern North America, human activities have favoured Acer saccharurn (VAM) over Fagus grandifolia (ECM). Girard and Fortin (1985) observed that coniferous forests in Quebec occurred on sites with soils that were less fertile, more acidic and had much slower organic matter decomposition rates, relative to sites with similar climates where VAM trees were also present. Most tropical plants have VAM associations, but areas where ECM trees are locally dominant occur in tropical forests throughout the world (Appendix 1). These ECM forests are warm throughout the year, but in some cases are subject to periodic drought (in Africa), or flooding (in South America). In these forests, trees with ECM are not randomly distributed but tend to occur in low-diversity stands or “groves” where VAM trees are excluded (Hogberg, 1986; Newbery et a l . , 1988; Singer and Araujo, 1979). Hart et a l . (1989) considered reasons for the local dominance of Gilbertiodendron deweveri (which has been reported to have ECM) in tropical forests in Zaire. These reasons were complex, but included the large seeds with limited dispersal and shade tolerance of seedlings of G. deweveri, which apparently were related to a gradual increase in dominance by this species during long periods without disturbance. Trees that form low-diversity stands in other tropical regions, often have similar reproductive characteristics (Hart et al.,

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1989) and in many cases represent islands of ECM activity surrounded by larger regions where trees with VAM are dominant (Malloch et a l . , 1980). In temperate regions forests dominated by ECM trees also generally have lower plant diversity than those containing trees with VAM (Berliner and Torrey, 1989; Malloch et al., 1980). Hogberg (1986) reported that soil phosphate levels were lower in ECM forests than in VAM communities in tropical Africa, but Newbery et al. (1988) found that seasonal variations in nutrient levels obscured these trends. Semi-arid African forests are often dominated by leguminous trees in the families Papilionaceae and Caesalpinaceae, which have either ECM or both VAM and nitrogen fixing Rhizobium associations, suggesting that conservation of soil nitrogen by tight nutrient cycling may reduce the benefit of nitrogen fixing associations in ECM communities (Hogberg, 1986). Ammonium and organic forms of nitrogen, are considered to be more restricted in supply that phosphorus in ECM forests, while phosphorus and nitrate may be more important limiting factors in VAM communities (Alexander, 1983; Girard and Fortin, 1985; Read, 1983). In the Amazon basin of South America, ECM trees occur in Igapo forests-areas where periodic inundation by “black water” rivers occurs, but trees in VArzea forests- where “white water” rivers cause flooding, have VAM (Singer, 1988). “Black water” rivers carry water which is nutrient poor and acidic relative to ‘*white water” rivers (Kubitzki, 1989). In this case there appears to be a good correlation between soil quality (low-nutrient supply, or acidity) and ECM tree dominance. It has been suggested that forests of trees with ECM can “degrade” soils by causing increased soil acidification (Harley, 1989). However, these trees may also be more likely to occur on soils that are naturally acidic, because they have a greater tolerance to low soil p H and high levels of aluminium or other toxic ions (Hogberg, 1986). Litter characteristics of ECM trees may be responsible for changes to soil properties, such as slow nutrient cycling, soil acidification and adverse allelopathic influences on other species (Section III.F.3.a). Leaf litter accumulation, as apposed to rapid mineralization or consumption of leaves by soil biota, is a distinguishing feature of both tropical and temperate ECM forests (Chabot and Hicks, 1982; Singer and Araujo, 1979). In ECM-dominated boreal forests nutrient cycling occurs slowly because decomposition is inhibited by the high lignin and tannin content of leaf litter (Horner et al., 1988). Litter decomposition may be further inhibited by low soil pH, or the withdrawal of water and nutrients by mycorrhizal fungi and roots (Section 1II.F.1). The activity and diversity of litter decomposing fungi can be substantially lower in ECM forests than in VAM forests (Singer and Araujo, 1979). Earthworms play an

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important role in decomposition processes in forests with VAM but are absent in boreal forests (Girard and Fortin, 1985) and termites may also be important in tropical forests with VAM. There is evidence that substantial differences in decomposition processes, which regulate the form and availability of nutrients, occur between ECM and VAM communities, but how this relates to the mycorrhizal strategies of dominant trees is uncertain. Estimates of carbon cycling in mycorrhizal associations support the contention that ECM associations are more expensive to maintain than VAM associations (Section III.E.7), originally proposed because of the higher fungal biomass associated with ECM roots (Harley, 1989, Harley and Smith, 1983; Read, 1983). Thus ECM associations require more investment in energy from the host than VAM associations, but in regions with short growing seasons ECM may be more advantageous if the function as perennial storage organs (Harley and Smith, 1983; Hogberg, 1986). Limited seed dispersal by ECM trees in tropical forests (Hart et a l . , 1989; Newbery et al., 1988) and support of seedlings by a pre-existing hyphal network (Newman, 1988) may also favour seedling establishment close to trees with the same type of mycorrhizas. It appears that ECM forests are more likely to occur in regions with cool climates or nutrient-poor acidic soils that limit plant productivity. However, there is no evidence that trees with ECM associations in temperate regions are inherently less productive than those with VAM associations. In tropical regions ECM trees may be marginally more efficient than VAM trees, since groves of ECM trees gradually expand into VAM forests during long periods without disturbance. It would appear that host leaf-litter characteristics, substrate utilization by mycorrhizal fungi, soil pH, nutrient levels etc. and environmental factors are the most important factors correlated with the mycorrhizal strategies of dominant trees, but their relative importance has not been established. The best way to summarize the information presented in this section is to present the following list of generalizations. (i) Most of the plant species present in natural communities throughout the world normally have mycorrhizal associations. (ii) On a worldwide basis, plants with VAM are predominant, ECM relationships are very common (and tend to dominate where they occur), while other association types (ericoid, orchid, etc.) are present in most ecosystems, but usually form only a small part of the community. (iii) Non-mycorrhizal or facultatively mycorrhizal plants are more likely to occur in habitats where (a) there has been severe, recent disturbance, (b) soils are very dry and/or saline and/or wet, (c) low temperatures prevail, or (d) soil fertility is abnormally high or extremely low. (iv) Dominant trees in a community may have either ECM or VAM associations, less often trees with both occur

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

257

together, but trees rarely if ever are non-mycorrhizal. (v) Throughout the world, plants in similar habitats usually have similar mycorrhizal strategies. (vi) Plants within a family usually have similar mycorrhizal relationships and plants within a genus nearly always do. We can be fairly confident that these generalizations would be reliable when used to predict mycorrhizal relationships at the community or ecosystem level, but they obviously should not be used to predict the mycorrhizal status of individual plants.

IV. CONCLUSIONS There has been considerable progress in our understanding of the occurrence and multifaceted role of mycorrhizas in ecosystems during the last 100 years, but despite increasing interest there is still much to be learned. It is now well established that the most plants in ecosystems have mycorrhizas, so studies of nutrient uptake, soil resource competition, nutrient cycling etc. in ecosystems may be of little value if they do not consider the role of these associations. The value of mycorrhizal associations to plants which occur in natural ecosystems, has been demonstrated for a limited number of species by growth experiments at realistic nutrient levels (Table 6). However, mycorrhizas provided little benefit to host plants in some of these experiments and attempts to demonstrate their value in natural communities (using fungicides which inhibit mycorrhizal activity) have largely been unsuccessful (Section III.E.7). While it is possible successfully to manipulate mycorrhizal associations in experiments using controlled conditions in a glasshouse or growth chamber, there may be difficulties when extrapolating these results to natural ecosystems. In particular, it is difficult to (i) find a substitute for undisturbed soil, (ii) remove mycorrhizal fungi from soils without substantially altering other biotic and abiotic soil properties and (iii) grow control plants without mycorrhizas without changing many aspects of their physiology. Experimental systems that have been used include (i) axenic culture of mycorrhizal fungi, (ii) mycorrhizal synthesis experiments using sterilized soil or axenic conditions, (iii) microcosm experiments, (iv) experiments using soil from ecosystems and (v) experiments in ecosystemsarranged in increasing order of complexity and predictive ability. The axenic culture of mycorrhizal fungi allows the influence of a single factor such as temperature to be tested on a range of endophytes, but some fungi cannot be grown in this way and in other cases results can be misleading (Section III.B.5). Mycorrhizal synthesis experiments are

258

M . BKCINDKET’T

more laborious but can allow the influence of one factor on a range of host-fungus combinations to be considered while environmental conditions are carefully controlled. Microcosm experiments can be used to study the influence of additional microbes or host plants in mycorrhizal experiments. These experiments have been used to study mycorrhizal influences on interplant nutrient transfer, competition between plants and decomposition processes (Dighton et a l . , 1987; Finlay, 1989; Grime et al., 1987: Perry et a l . , 1989). The interpretation of data from mycorrhizal synthesis experiments is complicated by the absence of complicating factors that occur in nature (Fig. 5.A) and the interactions between root properties, soil nutrient availability and mycorrhizal fungus activity which determine mycorrhizal responses in experiments (Fig. 5.B). Intact cores of unmodified soil from natural ecosystems can be used to study mycorrhizal establishment while maintaining control of environmental factors (Jasper et a l . , 1989c; Scheltema er a l . , l985a) and it is also possible to transplant seedlings into ecosystems and measure mycorrhizal formation (McAfee and Fortin, 1986, 1989; McGee, 1989: McGonigle and Fitter, 1988a). Most mycorrhizal experiments have been conducted using simplified systems (monocultures, dual-organism cultures, or sterilized soils) because it is expected that more complex systems would produce results that were highly variable or difficult to interpret. However, real soil conditions must still be kept considered when interpreting results of these experiments. For example, soil sterilization can create toxic conditions or increase nutrient levels (Sparling and Tinker, 1978b) and may enhance plant growth by removing pathogens (Afek et a!., 1990). The relatively small volume of soil available to plants in pots can reduce mycorrhizal benefits (Baath and Hayman, 1984). Additional complicating factors that may be important in ecosystems, but are usually excluded from experimental systems include preferences by endophytes for certain hosts or soil conditions which may occur in nature (Section 1II.D.l ) , soil organisms that may enhance the growth o r germination of (Section III.E.3), or are antagonistic to mycorrhizal fungi (Section III.B.3) and the impact of mixing or storing soil on mycorrhizal fungi (Section III.B.4). It should also be noted that pre-existing hyphal networks which may influence nutrient completion or reduce association costs are absent from most experiments (Section III.E.2) and variations in the mycorrhizal dependency of plants can mask competition effects in experiments (Newman, 1988). It is easy to criticise experiments conducted under controlled conditions because of their artificial nature, but results from experiments in natural ecosystems may not produce clear cut results if treatment effects are overwhelmed by other sources of variability (Allen and Allen, 1986; Allen et a l . , 1989b).

MYCORRHIZAS IN N A T U R A L ECOSYSTEMS

259

Measured responses of plants to mycorrhizal fungi will largely depend on soil properties which regulate nutrient availability (Section III.E.4), host plant nutrient requirements (Section III.E.7) and root system characteristics which determine nutrient uptake efficiency (Table 5 ) . These mycorrhizal responses must be measured by comparing improvements in plant growth, stress tolerance, etc. with those in a non-mycorrhizal control plant. However, most plants in natural ecosystems are normally mycorrhizal, so we are really measuring the magnitude of growth depression resulting from mineral deficiency in the non-mycorrhizal control plants and this will depend on the factors listed above (see Fig. 5b). When quantifying mycorrhizal benefits, it is best to construct response curves that compare the nutrient use efficiency of mycorrhizal and non-mycorrhizal plants over a wide range of nutrient levels (Abbott and Robson, 1984a, 1990b), or compare mycorrhizal and non-mycorrhizal plants with similar relative growth rates and phosphorus contents by manipulating soil phosphorus levels (Auge, 1989; Baas et al., 1989b; Graham, 1987; Pacovsky, 1986). Facultatively mycorrhizal plants should make the best experimental subjects, because these species can occur naturally without mycorrhizas. The existence of non-nutritional physiological differences between mycorrhizal and non-mycorrhizal control plants has not been properly established (Section III.E.6), but if these would be more likely to occur in obligately mycorrhizal species which are not found without these associations in nature. Roots in ecosystems have a much greater diversity in morphological features than is usually considered and this diversity involves features with the potential to influence water uptake, nutrient absorption and mycorrhizal formation (Sections III.C.l, III.E.4). For example, the mycorrhizal dependency of plants is influenced by root surface area and activity; long-lived roots often have suberized or lignified peripheral layers, which may enhance their survival when exposed to desiccation, pathogens etc., but should also restrict nutrient absorption; mycorrhizal associations may be regulated by root anatomical or chemical features; and root phenology can also influence the mycorrhizal relations of plants (Table 5 ) . Many mycorrhizal morphology studies would benefit from a greater understanding of root phenology and structure and ultrastructural investigations should include a preliminary survey of root features using histochemical staining procedures (Brundrett er a l . , 1990). It is likely that root systems have evolved in response to the plants need to obtain adequate soil resources (water and nutrients) while minimizing carbon costs, but optimal root system features for obtaining different resources may be in conflict (Section III.F.2). Correlations between root structure and mycorrhizal formation (Section III.E.4-5) suggest that the regulation of mycorrhizal associations as well as the

260

M . BRUNDRETT

need for efficient water and nutrient acquisition have influenced the evolution of root form. Agricultural selection may have also resulted in changes to root parameters that influence nutrient uptake and mycorrhizal formation (Section III.E.5; O’Toole and Bland, 1987). We are just beginning to understand the attributes of clonal isolates of mycorrhizal fungi that would result in the greatest benefit to associated plants and how they interact with environmental factors or soil conditions (Section III.B.5). In the future, careful identification of mycorrhizal isolates and progress in mycorrhizal taxonomy (especially with VAM fungi) may well reveal a much higher degree of fungus specialization with regard to these conditions. For this reason, it is advisable to assign isolate numbers to endophytes used in experiments, keep herbarium specimens and accurate information about where fungi were isolated so that future knowledge about taxonomic relationships and edaphic or climatic interactions can be taken into account (Morton, 1988, 1990; Trappe and Moline, 1986; Walker, 1988). Careful taxonomic studies and investigations of mycorrhizas in undisturbed ecosystems, should be considered to be essential foundations which ultimately benefit all mycorrhizal research. It could also be argued that we have much to learn about the ecology of mycorrhizal fungi in the natural ecosystems and soils where they evolved. Unfortunately, it is much easier to receive funding for research with more practical objectives involving domesticated plants or disturbed ecosystems. There are many questions involving the ecology of roots and mycorrhizal ecology that cannot be adequately answered at this time. There is evidence that the evolution of root form is regulated by trade-offs between features required for efficient direct nutrient absorption and water uptake on one hand and efficient mycorrhiza formation on the other, while minimizing root system costs (Section III.E.5). Roots exhibit less structural diversity than shoots. Has the need to form associations with a limited group of fungi (especially in the case of VAM) placed restrictions on the chemical and morphological evolution of roots? Have roots of non-mycorrhizal plants diverged more substantially (at least chemically) because this constraint on selection has been removed? Are processes such as rhizosphere modification, in addition to extensive root systems, used to help extract soil nutrients by non-mycorrhizal plants, such as members of Cruciferae, Cyperaceae and Proteaceae? Plants have evolved a very wide range of secondary metabolites and these chemicals appear to be more highly evolved in many nonmycorrhizal plant families (Table 4). Do these chemicals play a role in preventing mycorrhizal fungus establishment within the roots of nonmycorrhizal species? Are non-mycorrhizal more likely to have adverse allelopathic effects on mycorrhizal plants or their associated fungi? Do

MYCORRHIZAS IN NATURAL ECOSYSTEMS

261

mycorrhizal associations with local “edaphotypes” of mycorrhizal fungi help roots adjust to local soil conditions? Do associations with well adapted mycorrhizal fungi increase host tolerance to pollution, disease, etc., or are these benefits entirely due to increased mineral nutrient supply? How rapidly (and how) do mycorrhizal fungi adapt to soil or environmental conditions? How important are organisms o r processes which act as dispersal agents and consumers of mycorrhizal fungi in natural ecosystems? The role of mycorrhizal relationships in competition for soil resources requires further investigation. Nutrient cycling in ecosystems involves a large and complex web of saprobic organisms which usually culminates in mycorrhizally mediated uptake by roots. Thus nutrient uptake likely involves co-operation between mycorrhizal fungi and other members of this food web to obtain nutrients which are spatially or chronologically separated from roots. There is increased evidence that mycorrhizal fungi, especially those forming ECM or ericoid associations, can obtain nutrients form some (relatively simple) organic substrates that are usually not considered to be available to plants. Interplant nutrient transfer through a common mycorrhizal mycelium has been demonstrated experimentally (Section 1II.F. 1). Most evidence suggests that this transfer is not large enough to influence plant population dynamics, but some achlorophyllous plants live entirely by this means. There are a number of examples where interactions involving mycorrhizal associations appear to be important in ecological interactions between plants. Different forms of nutrient competition may occur between plants with different mycorrhizal strategies (Section III.F.2). Mycorrhizas may ameliorate plant competition by increasing the functional similarity of roots (Section III.F.2). Plant succession may involve changes to dominant mycorrhizal strategies or populations of mycorrhizal fungi (Section III.F.4). Plants with a particular type of mycorrhizal associations can be dominant in some areas and subservient or absent in others, apparently as a result of environmental o r edaphic factors (Section III.F.5). Some vegetation communities may have a tendency to be self-perpetuating because plants with other mycorrhizal types are inhibited by allelopathic interactions or soil property changes which they cause (Sections III.F.3.a, III.F.5). There is much scope for future ecological research which considers the role of mycorrhizal associations. The occurrence of mycorrhizas in ecosystems has been the subject of numerous investigations which have demonstrated their worldwide importance (Appendix 1). However, there is need for more research (i) in ecosystems where sampling has been limited and which allows a range of geographic locations within an ecosystem to be compared, (ii) which

262

M. RRUNDRE'M

considers the relative abundance of plants examined-so that the importance of mycorrhizal strategies can be determined at the community level (St. John and Coleman, 1983), (iii) which measures changes in mycorrhizal relationships or taxa along environmental gradients, such as soil moisture levels (Anderson et al., 1984; Ebbers et al., 1987; Lodge, 1989) or temperature (Koske, 1987a), (iv) which involves in silu identification of endophytes by mycorrhizal morphology or other means (McGee, 1989) and (v) which incorporates an understanding of root phenology and uses rigorously applied definitions to distinguish between different types or degrees of mycorrhizal associations. These types of studies may appear to be of little immediate practical value, but will inevitably be of great benefit to research involving forestry, revegetation etc. and will also help us understand the functioning of roots and mycorrhizas in agricultural situations.

V. APPENDIX 1 The following table contains a summary of our knowledge of the mycorrhizal relations of plants in natural ecosystems. Ecosystems are classified following Walter (1979). Only references in which roots were microscopically examined to confirm the presence and type of mycorrhizae are used in the table. If many references on mycorrhiza in an ecosystem were available, preference has been given to the most recent or concise surveys. The numbers of plant species sampled is provided for surveys of the latter type. Mycorrhizal studies which use the relative abundance of species examined to determine mycorrhizal colonization levels in the community as a whole are rare (St John and Coleman, 1983), but in most ecosystems a good appreciation of the overall importance of mycorrhizal can be obtained by combining floristic data (Hicks and Chabot, 1985; Larsen, 1980; Takhtajan, 1986) with mycorrhizal information from a number of surveys. N o attempt has been made to distinguish between facultative (poorly developed or inconsistent) and highly developed mycorrhizae. This table should only be used to obtain an overall picture of the worldwide importance of mycorrhizae and general changes in mycorrhizal strategies resulting from climatic o r edaphic trends, since the available data is often insufficient to allow accurate predictions of mycorrhizal relations in vegetation communities.

The Occurrence of mycorrhizae in natural ecosystems Ecosystem type Survey data: Location

n

Vegetation surveyed

Proportion of species with types of mycorrhizae

References

A. Evergreen rain forests

Africa, Asia. South America

trees and herbs: limited surveys

Southeast Asia

684

93% VAM, 7% NM

Janos (1987) (review)

trees: Dipterocarpaceae

ECM

Smits (1983), Alexander (1987). Lee and Lim (1989)

Africa

trees: most areas trees: locally dominant

VAM ECM

Hogberg (1986) (review). Newbery et al. (1988), Hogberg and Piearce (1986)

South America

trees in unflooded forests trees in seasonally flooded forests

VAM ECM

Singer and Araujo (1979)

ECM most VAM ERC

Chilvers and Pryor (1965), Warcup (1980) Langkamp and Dalling (1982) Reed (1987)

70% VAM, 30% ECM, 10% NM (ORC, ERC)

McGee (1986)

VAM ECM

Hogberg (1986) (review) Hogberg and Piearce (1986)

B. Seasonally dry tropical and subtropical forests Australia

Africa

dominant trees: Myrtaceae shrubs, herbs shrubs: Ericales

30

herbs. shrubs

93

most forest trees locally dominant trees

Continued

The Occurrence of mycorrhizae in natural ecosystems Continued Ecosystem type Survey data: Location

n

Vegetation surveyed

Proportion of species with types of mycorrhizae

References

C. Savanna Kenya

common grasses

5

VAM

Newman et al. (1986)

Cuba

trees

6

VAM

Herrera and Ferrer (1980)

Africa

trees - “Miombo” woodlands most other savanna trees

70% ECM VAM

Hogberg (1986) (review)

South America

herbs, shrubs: “Cerrado” vegetation

most VAM, also ORC, ECM

Thomazini (1973)

most have VAM NM plants usually also common

Bethlenfalvay et al. (1984), Bloss (1985), Bloss and Walker (1987), Hogberg and Piearce (1986), Khan (1974), Mejstrik and Cudlin (1983), Mukerji and Kapoor ( 1986), Rose ( 1980), Schmidt and Scow (1986)

most VAM, NM, ERC & ECM also common ECM ORC

Lamont (1982) (review) Chilvers and Pryor (1965), Gardner and Malajczuk (1988) Ramsay et a l . (1986)

55% VAM, 17% ECM &

M. Brundrett

VAM, 27% NM, ORC, ERC also

(unpublished data)

D. Hot deserts and semideserts North America, Galapagosherbs and shrubs Islands, Zambia, Algeria, India, Pakistan

E. Mediterranean regions Australia

Western Australia

shrubs, herbs trees: Myrtaceae herbs: Orchidaceae shrubs, herbs, trees: Eucalyptus forest

144 110

South Africa

shrubs, herbs

42% VAM, 58% NM

Berliner et al. (1989)

Europe

trees

most ECM

Meyer (1973) (review)

California

shrubs

VAM or ECM

Kummerow (1981)

41

F. Temperate broadleaved evergreen forests Northern hemisphere

trees: mostly Quercus spp.

ECM

Harley and Smith (1983), Meyer (1973)

Southern hemisphere

trees: mostly Nothofagus and Eucalyptus

ECM

Meyer (1973)

New Zealand

herbs: limited surveys

most VAM

Baylis (1967), Johnson (1977)

Japan

herbs: large surveys

most VAM

Asai (1934), Maeda (1954)

VAM or ECM

Brundrett et al. (1990), Kelly (1950), Kormanik (1981) Girard (1985), Lohman (1927), McDougall and Liebtag (1928)

G. Temperate deciduous forests North America

trees: dominant species well studied herbs (understorey): many surveys

most VAM, also NM 8z ERC

Ontario, Canada

trees, herbs, shrubs

68

80% VAM, 4% ECM, 13% NM, 3% ORC

Brundrett and Kendrick (1988)

Massachusetts, USA

trees, herbs, shrubs

45

71% VAM, 22% ECM, also NM, ERC & MON

Berliner and Torrey (1989)

Europe

trees: dominant species well studied herbs (understorey); many surveys

most ECM, some VAM

Le Tacon et al. (1987), Meyer (1973), Harley and Harley (1987), Dominik (1957), Dominik and Pachlewski (1956), Gallaud (1905), Real et al. (1976)

Europe FRG

beech forest herbs, trees

most VAM

28

79% VAM, 18% NM, 4% ECM

Mayr and Godoy (1989) Continued

The occurrence of mycorrhizae in natural ecosystems Continued Ecosystem type Survey data: Location

Vegetation surveyed

n

Proportion of species with types of mycorrhizae

References

H. Temperate grasslands most VAM

North America

grasses and forbs

Alberta, Canada

grasses and forbs

85

95% VAM, 5% NM

Currah and Van Dyk (1986)

California, USA

herbs: serpentine grasslands

27

98% VAM

Hopkins (1986)

Russia

herbs, forbs: steppe vegetation

68

88% VAM. 12% NM

Selivanov and Eleusenova (1974)

New Zealand

tussock grasses

VAM

Crush (1973a)

Europe, North America

herbs: semi-natural grasslands

most VAM

Gay et al. (1982), Read et al (1976), Sparling and Tinker (1978a), Rabatin (1979) Reeves et al. (1979)

Miller (1987) (review), Anderson et al . (1984). Davidson and Christensen (1977)

I. Cold deserts and semideserts Colorada, USA

herbs, shrubs: undisturbed site disturbed site

42

95% VAM, 5% NM

21

71% NM, 29% VAM

Wyoming, USA

herbs: undisturbed site disturbed site

22 16

95% VAM, 5% NM NM

Miller (1979)

Utah. USA

herbs: disturbed sites

74

57% VAM, 43% NM

Pendleton and Smith (1983)

Russia, Kazakistan

herbs, shrubs

234

65% VAM, 35% NM

Selivanov and Eleusenova (1974)

Worldwide

herbs, shrubs, trees

263

72% VAM, 26% NM, 2% ECM

Trappe (1981) (review)

Brundrett er al. (1990), Harley and Harley (1987), Le Tacon et al. (1987), Meyer (1973)

J. Boreal coniferous forests Northern hemisphere

trees: many studies

ECM

North America

shrubs, herbs: understorey

many VAM, also ERC, Girard (1985), Malloch and ECM, ARB, O R C & MON Malloch (1981, 1982)

British Columbia

herbs

Europe

shrubs, herbs: understorey

many VAM, also ERC, Dominik er al. (1954b), Harley ECM, ARB, O R C & MON and Harley (1987)

Northern hemisphere

herbs, shrubs

VAM, NM, ERC. ECM

Alaska, USA

dominant tundra plants

Canada, Ellesmere Is.

high arctic plants

24

54% NM, 25% ECM, 4% VAM, 4% ARB

Kohn and Stasovski (1990)

Sub-Antarctic islands

herbs

24

most VAM

Smith and Newton (1986)

Continental Antarctica

herbs

2

NM

Christie and Nicolson (1983)

27

78% VAM, 22% NM

Berch et a1 (1988)

K. Arctic vegetation

16

Katenin (1965). Laursen and Chmielewski (1980), Likins and Antibus (1982), Stutz (1972)

138% ECM, 31% ERC, 19% Miller (1982b) NM, 6% VAM, 6% ARB

Continued

The occurrence of mycorrhizae in natural ecosystems Continued Ecosystem type

Survey data: Location

Vegetation surveyed

n

Proportion of species with types of mycorrhizae

References

L. Alpine vegetation Europe, Asia, North and South America

Canada

trees: subalpine boreal forests

ECM

shrubs: above tree line herbs: many surveyed

ECM, ERC many VAM, ERC, NM & DSF also common

herbs, shrubs: alpine and montane

44

Dominik et al. (1954b), Keke and Yu (1986), Meyer (1973), Trappe (1988), Singer and Morello (1960) Haselwandter and Read (1980), Lesica and Antibus (1985), Miller (1982b), Nespiak (1953), Read and Haselwandter (1981)

16% VAM, 61% DSF, 9% NM, 14% ORC

Currah and Van Dyk (1986)

NM

Boullard (1939, Fries (1944) Rozema et al. (1986)

M. Edaphic vegetation communities 1. Wet Coastal Vegetation

Saltmarshes in Europe

dominant herbs: Juncus, Spartina, etc. herbs

18

56% VAM, 44% NM

Saltmarsh, USA

herbs

7

71% VAM, 29% NM

Cooke and Lefor (1990)

Saltmarsh, India

herbs

4

VAM

Sengupta and Chaudhuri (1990)

Mangrove vegetation

trees, herbs

Seagrasses

herbs: detailed root structure studies

26

NM

Lee and Baker (1973), Mohankumar and Madadevan (1986), Rose (1980)

NM (VAM not seen)

Cambridge and Kuo (1982). Kuo et a l . (1981), Tomlinson (1969) Anderson et a l . (1984), Bagyaraj et a l . (1979), Clayton and Bagyaraj (1984), Farmer (1985), Khan (1974), Read et a l . (1976), Sondergaard and Laegaard (1977)

2. Terrestrial Wetlands North America, Europe, Asia & New Zealand

herbs: hydrophytes in marshes. lakes, rivers

82

56% NM, 44% VAM

India

tropical hydrophytes

70

53% NM, 47% VAM

Ragupathy et a l . (1990)

Europe

herbs: wet meadows herbs, shrubs: peatlands.

73 40

86% VAM, 14% NM 68% NM, 15% VAM, also ECM, ERC & O R C

Mejstrik (1965, 1972) Hoveler (1982)

Temperate regions

trees, shrubs

VAM or ECM

Brundrett et a l . (1990), Keeley (1980), Lodge (1989), Harley and Harley (1987), Marshall and Pattulo (1981)

herbs, shrubs: moderately saline soils highly saline soils

VAM or NM NM

4 . Dry Saline Soils North America & Asia

Ho (1987), Khan (1974), Pond et al. (1984) Trappe (1987), Kim and Weber (1985) Continued

The Occurrence of mycorrhizae in natural ecosystems Conrinued Ecosystem type

Survey data Location

Vegetation surveyed

n

Proportion of species with types of mycorrhizae

References

N. Ecosystems created by disturbance 1. Sand Dune Vegetation

Europe. Holland

1s

herbs, shrubs

VAM

Ernst et al. (1984)

Europe. UK

6

most VAM

Nicolson (1960)

Europe. Italy

7

86% VAM. 11% NM

Puppi and Riess (1987)

VAM

Koske (198%) (review), Rose (1988). Sylvia and Will (1988)

North America Australia. Heron Island

12

57% VAM,43% NM

Peterson er al. (1985)

Australia. New South Wales

41

88% VAM. 19% NM. 2% ERC

Logan et al. (1989)

Hawaii

31

71% VAM. 25% NM

Koske and Gemma (1990)

Ahhre~~iariortsw e d : herb

=

herbaceous plant. VAM

=

vesicular-arhuscuiar rnycorrhizae. EChI = ectornycorrhizae. NM = non-mycorrhizai. = arhutoid rnycorrhizae. MON = rnonotropoid mycorrhizae. DSF = dark septate

ERC = ericoid rnycorrhizae. ORC = orchid mycorrhizae. ARB fungus.

MYCORKHIZAS IN N A T U R A L ECOSYSTEMS

27 1

ACKNOWLEDGEMENTS Most of this review was written while the author was supported by a postdoctoral fellowship from the Natural Sciences and Engineering Research Council of Canada. I am especially gratefull to Lyn Abbott and Allan Robson for their support. Some parts of this review developed during my tenure as graduate student under the guidance of Bryce Kendrick and Larry Peterson. A . H. Fitter, Lyn Abbott, David Jasper provided comments on the manuscript. I would also like t o thank Jim Trappe for his generous hospitality and access to his extensive collection of rare mycorrhizal papers.

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111. The Development of Misconceptions about “Tropical Soil” A . The Enormous Challenge of Mapping Soils on the 5-Billion

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Hectare Tropical Landscape B . Interdisciplinary Miscommunication about Soils in the Tropics C . The Tower of Babel Effect of Too Many Taxonomies and Nomenclatures ........................... D . Emphasis on Factors Rather than on Effects of Soil Formation: The 1938 Soil Classification IV . Advances in Soil Taxonomy and the Creation of the World SoilMap ...................................... A . Soil Taxonomy: A New Scientific Paradigm . . . . . . . . . . . B. FAO/UNESCO Soil Map of the World V . Diversity of Soil Taxa in the Tropics .................... A . General Soil Taxonomic Variation in Tropical Africa. America and Asia ........................... B . Ferralsols: Modern Inheritors of the Latisol Concept C . Acrisols: the Underestimated Soil Order . . . . . . . . . . . . D . Lithosols. Arenosols and Luvisols: From Extremely Fertile to Infertile, 500 Million Hectares Each E . Regosols, Yermosols and Cambisols: Weakly Developed Soils F . The Other 1 Billion Hectares: Extreme Variation VI . How Much Area in the Tropics Is Covered by Oxisols?: Results of the First Soil Surveys of the Amazon Basin . . . . . . . . . . . . . . . A . Background ............................... B . Some Common Properties of Amazonian Soils . . . . . . . . C. The New Soils Map ofthe Brazilian Amazon . . . . . . . . . D . How Much Area in the Tropics Is Covered by Oxisols? . . E . A Realistic Concept of Soil Diversity in the Humid Tropics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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VII. Meso- and Local-scale Soil Variation in the Tropics . . . . . . . . . . VIII. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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I. INTRODUCTION AND OBJECTIVES Concepts of diversity are central to our understanding of tropical plant and animal species, communities, and ecosystems. The concept of diversity as “richness in species” (Whittaker, 1977) is also important to understanding soils in tropical environments. Soil diversity in the tropics is affected by an extremely wide variety of parent materials, climates, biotic interactions, landforms and geomorphic elements, and soil ages. Many of these ecological factors that affect soils vary more widely in the tropics than in the temperate zone (Sanchez, 1976; Richter et a l . , 1985), and tropical landscapes commonly contain soils that range in age from amongst the youngest (e.g. Entisols and Inceptisols) to the oldest (e.g. Ultisols and Oxisols). Conceptual models of soil formation (Simonson, 1959; Jenny, 1941, 1980) also predict that environments present in tropical regions result in extreme heterogeneity in soil-ecological processes, soil properties, and consequently soil taxa. Such soil heterogeneity is expressed on regional landscape to microsite scales (Drosdoff, 1978; Buol et al., 1989). Despite ample evidence for the heterogeneity of soils in the tropics, the popular concept of a “tropical soil” often suggests that between the tropics of Cancer and Capricorn, ihere are special groups of soils with specific properties and problems. Early European natural historians and scientists, impressed by the prolific forest vegetation of the humid tropics, often concluded that tropical soils were extremely rich and fertile. Alfred Russel Wallace (1853) in Travels on the Amazon and Rio Negro exclaimed about Amazonia, “For richness of vegetable productions and universal fertility of soil, it is unequalled on the globe ..... Partly in reaction to this early exuberance about the fertility of tropical soils, many subsequent observers have fervently claimed that after native vegetation is cleared, “tropical soil” was extremely fragile and incapable of sustaining land development and management. Soils that support tropical rainforests have been commonly described with words such as “deserts covered by trees” (Goodland and Irwin, 1975). In this paper we emphasize that soils in the tropics can no longer be generalized as being universally fertile or dangerously fragile; they can, however, be generalized by their remarkable diversity.

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“Tropical soil” is commonly considered to be a deep, red, friable soil that is acid, well drained, and lacking distinct horizons. This soil is often associated with the Oxisol soil order of the Soil Taxonomy system (Soil Survey Staff, 1975, 1978a; Buol and Eswaran, 1988). Oxisols are similar although by no means identical to Latosols in the 1938 USDA soil classification system (Baldwin et al., 1938; Kellogg, 1949; Cline, 1975), Latossolos of the Brazilian classification system (Carmargo et al., 1986), Ferralsols in the FAO/UNESCO (1974) system, and Ferrallitic soils in the French system (Duchaufour, 1982). Such soils are often inferred to dominate tropical landscapes. Laterization is said to control such soils, an old and persistent idea that suggests that the intense leaching of the humid tropics has depleted the typical soil profile of nearly all nutrients, destroyed all clay minerals by dedication, and thereby concentrated sparingly soluble metal oxides of iron and aluminium. Laterization has also been suggested potentially to turn soils into a hardened mass of laterite following clearing of original vegetation (McNeil, 1964). In the past, laterite was considered to be a soil type toward which nearly all “tropical soil” was evolving. Such ideas about soils in the tropics retain a curious but persistent currency, despite highly persuasive papers and books that describe the diversity of soil conditions in this large and heterogenous region (Sanchez and Buol, 1975; Sanchez, 1976; Drosdoff, 1978; Buol et al., 1989; Greenlands, 1981). Over the last century, concepts of tropical soils have tended to oversimplify their taxonomy (e.g., McNeil, 1964), and to emphasize a relatively straightforward relationship between regional climate and soil (Fig. 1). Despite this frequent oversimplification, the technical scientific literature about soils in the tropics has actually contained a variety of perspectives about soil formation and climate influences (Hilgard, 1906; Sibirtev, 1914; Joffe, 1936; Jenny, 1941; Kellogg, 1949; Mohr and Van Baren, 1954; Sanchez, 1976). Climate factors are obviously important to soil development, but so are parent materials, landforms, hydrology, biological organisms, and soil age. Popular notions about “tropical soil” typically diminish the complexity of the interactions among these environmental factors. Soils are highly dynamic biogeochemical systems, influenced by multiple interacting factors. We make no pretence that our critique of the concept of a “tropical soil” is a new idea. A long and perceptive series of publications emphasize the diversity of soils in the tropics. Such a perspective has been expressed regularly throughout the twentieth century. Hilgard’s (1906) classical soil-ecology treatise, Soils,described a variety of the soil conditions in the tropical world and emphasized the notable absence of reliable data about soils in the tropics. Laterite has often been viewed as the representative “tropical soil” (Sibirtzev, 1914), but even Hilgard

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(1906) was puzzled by claims that laterite soils were supposed to be the most characteristic soils of the tropics. In the 1930s, Hardy (1933) argued that laterite occupied a relatively minor area of the tropics, and based on South American and African experience, De Camargo and Vageler (1937) vigorously contended that despite a so-called “logical conclusion” that laterite should occur throughout intertropical regions, there was nearly an “absolute absence of real laterite in the whole region.” De Camargo and Vageler (1937) criticized soil maps that were based on extrapolations from broad patterns of temperature, rainfall, and geology, which they claimed overestimated the occurrence of laterite, and also underestimated the content of soil organic matter. In the 1940s, Pendleton and Sharasuvana (1942, 1946) also criticized generalizations made about so-called “laterite tropical soils”, especially those descriptions about laterite and latosols “found in leading works on soils”. Some more recent and notable publications that describe many aspects of tropical soil diversity include Nye (1954), Nye and Greenland (1960), Travernier and Sys (1965), Sombroek (1966), D’Hoore (1968), Buol (1973), FAO-UNESCO (1974, 1988), Bornemisza and Alvarado (1974), Sanchez and Buol (1975), Soil Survey Staff (1975, 1986, 1987a, b), Sanchez (1976), Lepsch et al. (1977), Van Wambeke (1978), Drosdoff (1978), Cole and Johnson (1978), Perez et al. (1979), Greenland (1981), El-Swaify er a l . (1982), Bleeker (1983), Burnham (1984), Cochrane et al. (1985), Lathwell and Grove (1986), Vitousek and Sanford (1986), La1 (1987), and Buol et al. (1989). Although there have been many soil surveys and major conceptual developments since Hilgard’s time (1906), similar observations can still be made today; that there is a diversity of soil conditions in the tropics, yet a notable absence of reliable soil data in many tropical regions. Many landscapes of tropical America and Asia have now been systematically surveyed at a relatively small mapping scale (<1:1 million), yet major concentrations of unsurveyed soils in the tropics remain, especially in tropical Africa. In recent decades, soil science has lost many of its parochial perspectives and temperate-zone biases. Increasing numbers of soil scientists and ecologists from tropical nations have provided substantial pressure for making soil concepts more realistically global in application. Nonetheless, notable similarities exist between the criticisms of “tropical soil” in the 1930s and 1940s (Hardy, 1933; De Camargo and Vageler, 1937: Pendleton and Sharasuvana, 1942, 1946) and those made in more recent years (Sanchez, 1976; Drosdoff 1978; Buol et a l . , 1989). We are impressed by the persistence of misconceptions about “tropical soil”, but also with the rapid scientific developments made in recent years. Even a book such as Soils of the Humid Tropics (National Academy of Science, 1972), an excellent “summary of present knowledge” written in

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the early 1970s by widely respected scientists, now gives the reader a clear sense of how rapidly soil science and ecology of tropical ecosystems have progressed in recent years. The objectives of this paper are to evaluate common misconceptions about “tropical soil” that are of ecological significance, to consider reasons why such misconceptions have persisted over many years, and to evaluate quantitatively a hypothesis originally made by Sanchez and Buol (1975) that soils in the tropics are noted for their marked taxonomic diversity. Heterogeneity of soils throughout the tropics is quantified by estimating areal extents of soil taxa from recent soil maps of tropical America, Africa, and Asia. Soil diversity in the humid tropics is evaluated by examining the first comprehensive soil maps of the Brazilian Amazon River basin, a 500 million hectare region where systematic soil surveys have only recently been completed. Geographic Information Systems (GIS) are used to compare the FAO/UNESCO (1971, 1974) and the more recent Brazilian soil maps of Amazonia (EMBRAPA, 1981; Carmargo et al., 1986).

11. CHANGING PERSPECTIVES ABOUT SOIL TAXONOMY In order to understand more fully soil diversity in the tropics, an understanding of soil taxonomy and classification is mandatory. Traditionally, soil taxonomy has been strongly guided by ideas about soil genesis and formation. In other words, environmental factors have traditionally been used to provide the structure of soil taxonomy. This soil taxonomic emphasis on the factors of soil formation is a distinctly nineteenth century Russian concept that has dominated soil and ecological sciences from the 1880s until at least the 1950s. Great soil groups (major soil types) such as Podzols, Chernozems, Tundra soils, and Latosols have been recognized by soil scientists and ecologists, from the time of Dokuschaev and his colleagues in Russia (Sibirtzev, 1914) to Whittaker (1975), the noted American ecosystem ecologist. Traditionally, most great soils groups were associated with major bioclimatic regions. Latosols, a soil considered to exist almost entirely in the “torrid and humid tropics” (Sibirtzev, 1914), were often conceived as the typical soil type that forms under hot and humid tropical climates (Fig. 1). Although Dokuschaev is widely recognized for developing the concept of five basic soil-forming factors (climate, parent material, biology, topography, and soil age), the predominant factor actually emphasized by Dokuschaev was climate (Fig. 1). This emphasis on climate has had major implications for how soils in the tropics have been popularly perceived.

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Classifying soils using soil genetic concepts was often a deductive process since representative and reliable soils data were often lacking, especially in the tropics. The absence of reliable soil composition data for representative soils in the tropics has been consistently noted by many critical thinkers, at least as far in the past as Hilgard (1906). Even until the 1980s, for example, systematical soil surveys and soil maps of the Amazon basin (approximately 10% of the tropics as a whole) were entirely non-existent. Comprehensive soil surveys are still non-existent for soils in the Zaire River basin in central Africa. Many soil maps in the tropics have thus been based on general climatic, vegetative, and physiographic information, rather than on observations of the soils themselves. Major soil taxonomic problems were created because soil classification was initially based on inferred factors of soil formation rather than on empirical soil data. Since genesis of nearly all soils has occurred over long periods of time (i.e. in the past), many soil forming factors can only be inferred and are not evident for specific soils (Smith, 1983). For good reasons, the historical emphasis on soil genesis and formation has recently been radically altered by the publication and use of Soil Taxonomy (Soil Survey Staff, 1975; 1987b). The Soil Taxonomy system and a number of other modern classification systems, e.g. the Brazilian classification system (Carmargo et al., 1986, have directed attention to quantifiable soil data and away from speculative factors of soil formation. This change in perspective represents a fundamental change in the scientific paradigm of soil taxonomy, a change that is described well by Kuhn’s (1970) definition of scientific revolution. The change in perspective has had significant implications for understanding of soils in general (Smith, 1983), but especially for soils in the tropics. Since relatively little quantitative data have been available to describe soil conditions in the tropics, soils in the tropical environment have been subject to much misleading speculation. The quantitative approach of the Soil Taxonomy system has put priority on detailed soil surveys and descriptions such as the recently completed Brazilian soil surveys of the Amazon basin (EMBRAPA, 1981). The Soil Taxonomy system has forced soil and ecological scientists to collect empirical data on the soils themselves, to re-examine old ideas about soils in the tropics, and to continue to develop new soil concepts (Soil Survey Staff, 1986, 1987a, b; Buol and Eswaran, 1988). The new Soil Taxonomy system is organized into eleven soil orders, all of which are well represented in the tropics. Soil orders are briefly described in Table 1 and are concepts discussed throughout the paper. In summary, the popular notion of “tropical soil” developed directly from the traditional paradigm of a soil taxonomy that was based on soil-forming factors. The recent change in taxonomy to quantitative and

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Table 1 Eleven soil orders in soil taxonomy describe all of the earth’s soils; all eleven orders can be found in the tropics (Soil Survey Staff, 1987a,b; Leamy et al., 1988) Order

Brief description

Alfisol

Soils with clay-enriched subsoils (argillic or kandic) with moderate to high pH and base saturation (35%).

Andisol

Soils derived from volcanic ash with > 35 cm organic-rich surface soil horizon (andic).

Aridisol

Soils dry > 50% of the year with no organic rich surface horizon.

Entisol

Soils with little if any soil profile development.

Histosol Inceptisol

Soils with > 30% organic matter to a depth of > 40 cm. Soils with incipient soil profile development (more development than Entisol).

Mollisol

Soils with deep surface horizon with high organic matter and moderate to high pH and base saturation (mollic).

Oxisol

Soils with little texture difference between surface and subsoil horizons with low cation exchange capacity and high iron and aluminium oxide content (oxic).

Spodosol

Soils with subsoil horizon having high accumulations of organic matter, iron, and aluminium (spodic).

Ultisol

Soils with clay-enriched subsoils (argillic or kandic) with low pH and base saturation (35%).

Vertisol

Soils with > 30% clay in all horizons with large shrink-swell properties upon drying and wetting, respectively.

descriptive approaches marks a fundamental change that greatly impacts soil and ecological sciences. This change has yet to be fully incorporated into the discourse of soil scientists and ecologists. Our intention in this paper is to encourage the process of paradigm change in soil taxonomy, especially among soil and ecological scientists with interests in the tropics.

111. THE DEVELOPMENT OF MISCONCEPTIONS ABOUT “TROPICAL SOIL” Soil science developed rapidly in the second half of the nineteenth century, largely in temperate zone regions of Russia, western Europe,

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and North America. Traditional definitions and theoretical concepts about “what soil is” and “how soil develops” were derived almost entirely from this temperate-zone experience. Although many early studies of soils in the tropics to the warm temperate zone can be credited for their insights (e.g. Hilgard, 1860; Marbut and Manifold, 1926; Mohr, 1944), soil science and ecology benefited little from soil conditions in tropical environments. This early temperate-zone bias was a serious deficiency for the development of soil science and ecology, especially considering that the terrestrial tropics cover about 40% of the earth’s terrestrial surface. Ideas about tropical soils were for many years closely associated with laterite, a geologic material (not a soil) first described in southern India by Buchanan (1807). Buchanan (1807) described laterite as one of the most valuable materials for building. It is diffused in immense masses, without any appearance of stratification, and is placed over the granite that forihs the basis of Muluyulu. It is full of cavities and pores, and contains a very large quantity of iron in the form of red and yellow ochres. In the mass, while excluded from the air, it is soft, that any iron instrument readily cuts it, and is dug up in square masses with a pick-ax, and immediately cut into the shape wanted with a large trowel, or large knife. It is (sic) very soon after becomes hard as brick, and resists the air and water much better than any bricks that I have seen in India . . . (1)n several native dialects, it is called the brick-stone (Zticu cullu). Where, however, by washing away the soil, part of it has been exposed to the air, and has hardened into a rock, its color becomes black, and its pores and inequalities give it a kind of resemblance to the skin of a person affected with cutaneous disorders; hence in the Turnul language it is called Shuri cull, or itch stone. The most proper English name would be Laterite, from Lateritis, the appellation that may be given to it in science.

Buchanan obtained the term “laterite” from later, meaning “brick” in Latin. By the late nineteenth century, the term laterite was in general use throughout India with detailed descriptions recorded (Meldicott and Blanford, 1879). In fact, great interest in laterite was stimulated throughout the tropical world (Fox, 1923; Harrassowitz, 1926). By the end of the nineteenth century, laterite had also been identified in surficial formations in South America, Africa, and Australia (Sivarajasingham et al., 1962). Maps of laterite suggested very wide distribution (Fig. 2). In the popular mind, laterite became a soil type, and was no longer restricted to Buchanan’s (1807) narrow definition of a geologic material. Laterization became the soil process considered to be representative of “tropical soil,” and laterite became for many the material toward which soils in the tropics naturally evolved (Sibirtsev, 1914). These perceptions of “tropical soil” did not go uncontested. From the time of Buchahan’s first description, laterite created controversy among

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Fig. 2 Historic maps of Africa and India with areas of laterite suggested in black (after Prescott and Pendleton, 1952). Such maps reflect old notions of the wide extent of laterite across the tropics. Some authors (e.g. McNeil, 1964) have mapped laterite even more extensively than Prescott and Pendleton (1952).

soil scientists and geologists (Mohr and Van Baren, 1954; Sivarajasingham et al., 1962). Definitions of laterite ranged widely and little consistancy was followed. To many, all red soils in the tropics came to be considered as laterite or at least lateritic. Efforts were made to keep the definition of laterite as narrow as Buchanan’s, and to distinguish between laterite and lateritic on the basis of chemical composition. One example is that of Fermor (1911), who defined “laterite soils” as containing >90% oxides of Fe, Al, Ti, and Mn; “siliceous laterite soils” as containing between 50 and 90% metal oxides; and “lateritic soils” as containing 25-50% metal oxides. The terms “laterite” and “lateritic” received very wide use by soil and ecological scientists, following the

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adoption of these terms as names of soils in the 1938 soil taxonomic system of the US Department of Agriculture (Baldwin et al., 1938). In 1949, the concept “Latosol” was proposed by Kellogg (1949) as a major soil type and as an amendment to the 1938 USDA classification system. Kellogg’s (1949) intention for the term “Latosol” was to help avoid real and semantic problems that surrounded the terms laterite and lateritic. Kellogg (1949) very tentatively proposed the concept of Latosol, specifically because of his high regard for the taxonomic uncertainties presented by the diversity of soils in the tropics. Nonetheless, the Latosol proposal proved very controversial (Prescott and Pendleton, 1952; Mohr and Van Baren, 1954). Despite controversy and an absence of accepted definitions and terminology, a set of soil characteristics became closely associated with “tropical soil”, characteristics that are often synonymous with the process and products of laterization. These soil characteristics include: (1) extremely intense weathering and leaching; (2) low contents of soil organic matter; (3) destructive weathering of soil alumino-silicate clay minerals; (4) low nutrient retention capacity; (5) an ability to harden irreversibly upon exposure to sun and air; and (6) a homogeneity in physical and chemical properties, both horizontally over the landscape and vertically within the soil profile. These six soil characteristics have veracity when applied to certain soils found in the tropics, however, they cannot be used to generalize about the soil conditions found within the 5-billion hectare tropical landscape. Many soils in the humid tropics are, of course, highly leached and intensively weathered; some have relatively low contents of organic matter and alumino-silicate clay minerals, and thus low cation exchange capacity. Many highly weathered upland soils in the tropics have relatively little apparent distinctions between horizons, and are red and well drained. Relatively few soils, however, appear to have problems with irreversible hardening of laterite. Because these six characteristics are so often associated with “tropical soil”, each is briefly described and evaluated.

I . Intense Weathering and Leaching Over pedogenic time, soil leaching by intense and abundant rainfall at relatively high temperature removes relatively soluble soil constituents such as mineral nutrients and even silica. Leaching and dedication remove nutrients from soils, and concentrate the low solubility oxides of iron, aluminium, titanium, chromium, and nickel. The rate at which leaching and weathering can form Oxisols has been greatly overestimated in the past (Van Wambeke et af., 1983), an overestimate that developed mainly from ideas about soil formation rather than from data

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derived from soils themselves. Van Wambeke et al. (1983), for example, emphasize that because clay minerals of Oxisols (Table 1) are often dominated by kaolinite rather than gibbsite, the desilication of Oxisols even in the humid tropics may seldom reach the intensity required for gibbsite [Al(OH)3] domination of the soil clay fraction. In fact, although Oxisols represent soils in advanced stages of weathering, they are most commonly found in materials that have been transported and pre-weathered during earlier geologic periods (Buol et al., 1989). The materials that compose Oxisol soils may often have been subjected to several leaching and weathering cycles. Moreover, Oxisols that are not formed from transported, pre-weathered materials are typically derived from geologic materials that weather rapidly, e.g. basalts. In sum, although many soils in the tropics are weathered very intensively, soils in the tropics can best be generalized by their extremely wide range of weathering states. The pedogenic time and the intensity of weathering that are required to form Oxisols should not be underestimated.

2. Low Soil Organic Matter Accumulation of soil organic matter (SOM) results from the balance between carbon inputs and outputs to and from soil; i.e. between plant production and microbial respiration of plant products. Soil organic matter has long been suggested to be relatively low in “tropical soil”, due to the idea that at the relatively high temperatures of the tropics, microbial respiration and decomposition may be stimulated more than plant production (Jordan, 1985, 1988). However, there are relatively few data to support this contention, and the SOM of forested soils in the humid tropics may frequently exceed 16 kgC/m2 in the surface metre depth of soil (Post er a l . , 1982; Zinke et al., 1984), a substantial amount of soil C for mineral soils (Soil Survey Staff, 1975). The idea that soils of the tropics have relatively low SOM is criticized by a number of scientists in recent years (Anderson and Swift, 1984; Sanchez et al., 1982; Buol et al., 1989). In fact, Sanchez et al. (1982) indicate that SOM in Mollisols, Alfisols, and Ultisols (Table 1) in temperate climates differs little from comparable soils in the tropics, and that SOM in Oxisols of humid tropical climates is no lower than in similar textured soils on comparable landforms in temperate zones. Much is not understood concerning soil organic matter; for example, some Oxisols on upland landforms have very large acculmulations of SOM in deep, “humic” surface horizons, apparently indicating that soil clay minerals have stabilized and protected organic compounds against decomposition (I.F. Lepsch, personal communication). Detailed analyses of the very large soil data set assembled by Post et al. (1982) and updated in Zinke et al. (1984), reveal two points about

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SOM in soils of the tropics. First, SOM in soils is highly variable. Mean estimates of SOM shown in Fig. 3 for soils with > 24 "C and annual PET/PPT ratios > 0.5, have standard deviations that approach 100% of the mean, despite relatively large sample sizes ( n > 120). Second, there is little indication that SOM is lower in the tropics compared to ecosystems in the temperate zone that have comparable soil moisture regime, i.e., that have similar PET/PPT ratios in Fig. 3. Controlling for soil moisture regime, the largest and most comprehensive global soil collection indicates that SOM in the tropics appears similar in content to that in soils of subtropical, warm, and cool temperate climates. Although plant productivity and microbial respiration are strongly controlled by temperature and moisture regimes, the interactive effects of temperature and moisture on the balance between plant productivity

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and microbial respiration (i.e. SOM) need much better evaluation. The mediating effect of soil texture also needs consideration in these evaluations, because clay minerals often can adsorb and protect organic compounds from microbial decomposition.

3. Weathering of Alumino-Silicate Clay Minerals Contents of alumino-silicate clay minerals in some soils of the tropics are low, due to intense leaching and weathering conditions that eventually decompose weatherable clay minerals that are inherited from parent material (primary minerals) or that have formed during soil development (secondary minerals). Destruction of alumino-silicate clays is often emphasized to be nearly complete in “tropical soil”. The rate of weathering and decomposition of clay minerals has often been overestimated, as kaolinite is the most common clay mineral in many soils of the humid tropics, and is frequently observed to dominate clay mineralogy of Oxisols (Van Wambeke et a l . , 1983). Recently, kaolinite-dominated subsoils have received detailed research attention (Moormann, 1985), and such work has recently led to the recognition of a new diagnostic soil horizon in Soil Taxonomy (Soil Survey Staff, 1987a). These are kandic horizons (kaolin-dominated) and are relatively common in the humid tropics, sub-tropics, and temperate zones. Kandic horizons will hopefully make it easier to distinguish Alfisols and Ultisols (Table 1) from Oxisols (Moormann, 1985). As more soil mapping is accomplished, kandic horizons of Ultisols (Moormann, 1985) may well replace oxic horizons of Oxisols in many areas of the humid tropics, especially in Africa (C. Sys, personal communication).

4. Low Cation Exchange Capacity A soil’s cation exchange capacity (CEC) is one of its most important attributes for controlling nutrient availability to plants and the chemistry of drainage waters. Negative charge in soil (soil CEC) also gives an ability to retain positively charged nutrient cations against leaching, and to control plant-availability of potentially toxic cations such as aluminium. Soil CEC is actually the net negative charge possessed by soils, and under acidic conditions soils may develop additional positive charge. Soils with such properties are known as variable charge soils. Although variable charge is a soil property found in soils from boreal to tropical climates, many soils in the tropics have significant variable charge and under acidic conditions may have difficulty retaining base cation nutrients (Uehara and Gillman, 1981). Variable charge means that the electrical charge possessed by soils is not constant and that at low pH, a soil’s positive charge may reduce net cation retention. Although most

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soils in the field appear to have net negative charge even at low pH (Cochrane et af ., 1985), of potentially great significance, however, are Acric Oxisols (Buol and Eswaran, 1989), soils with clays that have CEC that approaches zero (< 15 mmol,/kg of clay). Low CEC is a problem of many soils throughout the world. Whether due to low soil clay mineral content, low soil organic matter, or acidified variable charged surfaces, low net CEC gives a soil little ability to retain nutrient cations against leaching. Low CEC is thus a biologically significant soil property, and sustained productivity of low CEC soils is dependent on careful soil nutrient management. Soil management principles and technologies have been developed for low CEC soils, especially in southeastern Asia, southeastern USA, and elsewhere, and long-term management of such soils is possible with attention paid to soil organic matter, soil chemistry and fertility, and agronomic factors.

5. Irreversible Soil Hardening The popular notion of “tropical soil” has often contained the idea that once a tropical soil is exposed to the elements, it will irreversibly harden. Much attention has been devoted to the supposed threat that such laterite soil poses for agriculatural development in the tropics (e.g. McNeil, 1964; Sanchez, 1976). Recent soil surveys demonstrate clearly that soils with laterite (which is now called plinthite) occur in a relatively small portion of the tropics, exactly as predicted by De Carmargo and Vageler (1937). This is actually a major change in thinking compared to claims about laterite’s distribution that were made in the past (Fig. 2). Plinthite nodules or layers are found in a variety of soil taxa, in Alfisols, Ultisols, and Oxisols, and actually appear to be more prevalent in Alfisols and Ultisols than in Oxisols (Van Wambeke et af., 1988). Iron pans and other indurated layers also occur in a variety of soils in the tropics. It is unfortunate that scientists and other observers have spent such attention on laterite hardening as a potential problem in the tropics. Given the other soil physical problems that have nothing to do with laterite, there is no question that soil physical properties are critical to sustaining the potential productivity of soils following the clearing of native vegetation (Lal, 1987). Important physical properties of soils that are susceptible to land use degradation (unless adequate vegetation is maintained) include soil structure, porosity, and infiltration capacity. As environmental problems, each is much more significant than the problem of laterite to the long-term management of soils in the tropics.

6. Homogeneity of Soil Properties Variability of soil properties depends on the property of interest and on

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areal and vertical scale factors (Wilding and Drees, 1983). Especially on ancient land surfaces or in old sedimentary materials, minerals may have been intensively weathered, and most nutrients long since released for plant use or to leaching. Extreme weathering tends to diminish inherited differences in soil materials in oxic horizons. However, the notion that “tropical soil” has properties that are somehow homogeneous, is closely associated with soil genetic ideas that leaching and weathering have been somehow complete in soil materials found in the humid tropics. Certainly the idea that tropical soil is homogeneous over wide areas did not achieve its prominance from soil variability data such as those collected by Lopes and Cox (1977) or Wilding and Drees (1983). Considering that the tropics cover 40% of the earth’s terrestrial surface, much detailed spatial variability data have yet to be collected. A lack of soil horizonation is also supposed to be characteristic of “tropical soil”. Certain types of Oxisols lack distinct soil horizons, and they appear to have relatively simple soil profiles: remarkably uniform with soil depth. Presumably such soils have lost their horizons during intensive leaching of past and present weathering environments. Generalization about an absence of distinct horizonation in Oxisols can be easily exaggerated, especially since many Oxisols actually have prominant, deep horizonation. Some Oxisols have large accumulations of organic matter in surface horizons (Sanchez et a l . , 1982), sometimes to a depth of a metre o r more, due to the organic compounds being stabilized by clays. Moreover, if soils include the entire soil-weathering profile, some Oxisols with well developed horizonation may total tens of metres in depth (Eswaran and Wong, 1978). Apparent similarities among Oxisols in the tropics are often in fact misleading. Physical and chemical properties of Oxisols are often inherited from parent materials, despite long-term intense weathering (Greenland, 1981). Oxisol soils vary considerably, as is described in Table 2 which includes the numerous and varied subtaxa that the Oxisol order now contains (Buol and Eswaran, 1988). The Oxisol order now has five suborders that are based on moisture regime (Table 2), from soils with high water tables and excess water (Aquox) to those that are Table 2 Description of new soil suborders and great groups of Oxisols that demonstrate the diversity of soils within the Oxisol order (Buol and Eswaran, 1988)

Suborders or great group

A. Aquox: B. Perox: C. Udox:

Brief description Water saturated within 30 cm > 30 daysfyear Perudic soil moisture regime (moist all year, but no lengthy saturation) Udic soil moisture regime (< 90 daysfyear too dry to plant) Continued

SOIL DIVERSITY IN THE TROPICS

33 1

Table 2 Continued Suborders or great group

D. Ustox: E. Torrox:

Brief description Ustic soil moisture regime (moist > 90 days, < 270 days/year) Aridic soil moisture regime (moist < 90 days/year)

A. Aquox: Most are small in area. Acraquox: Plinthaquox: Eutraquox: Haplaquox:

> 5 in ECEC < 1.5 mEq/100 g clay within 200 cm; PHKC~ oxic horizon Continuous plinthite within 125 cm > 35% BS @ pH 7 within 135 cm Other Aquox soils

B. Perox: Precipitation year. Sombriperox: Acroperox: Eutroperox: Kandiperox: Haploperox:

5

Potential evapotranspiration in all months of the

High SOM in sombric horizon < 150 cm ECEC < 1.5 mEq/100 g clay; PHKCL> 5 in oxic horizon > 35% BS @ pH 7 within 125 cm > 40% clay in surface 18 cm layer with kandic horizon < 150 cm of surface Other Perox soils

C. Udox: previously called Orthox, “true” Oxisols, in 1975 Soil Taxonomy. Sombriudox: Acrudox: Eutrudox: Kandiudox: Hapludox:

High SOM in sombric horizon < 150 em ECEC < 1.5 mEq/100 g clay; PHKCL> 5 in oxic horizon < 150cm > 35% BS @ pH 7 within 125 cm > 40% clay in surface 18 cm with kandic < 150 cm of surface Other Udox soils

D. Ustox: probably the most extensive Oxisol suborder. Sombriustox: Acrustox: Eutrustox: Kandiustox: Haplustox: E.

High SOM sombric horizon < 150 cm ECEC < 1.5 mEq/100 g clay; PHKCL> 5 in oxic horizon > 35% BS @ pH 7 within 125 cm > 40% clay in surface 18 cm with kandic < 150 cm of surface Other Ustox soils

Torrox: high B.S.; often excellent crop soils when irrigated.

Acrotorrox: Eutrotorrox: Haplotorrox:

ECEC < 1.5 mEq/100 g clay; PHKCL> 5 in oxic horizon > 35% BS @ pH 7 within 125 cm Other Torrox soils

ECEC is effective cation exchange capacity; pHKa is soil pH in 1 M KCI; BS is base saturation at soil pH of 7; and SOM is soil organic matter.

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D . D. RICHTER AND L. I. E A B E A R

generally dry (Torrox). Chemically and physically, Oxisols also vary widely (Table 2). Texture ranges from low to very high clay contents, and from relatively nutrient-rich (Eutro- great groups) and to those that have extremely low CEC (Acr- great groups). Conceptual problems often arise because the six characteristics discussed above are used to generalize about soils on regional scales in the tropics, far beyond any bases in observed fact. Popular misconceptions about soils in the tropics have persisted for at least four reason: (1) it has been an exceedingly complex task to characterize and map the soils on the 5-billion hectare tropical landscape; (2) scientists of different scientific disciplines have not communicated effectively; (3) no common soil taxonomy has ever existed that could classify soils and provide a common language for discussing soil problems throughout the tropical world; and (4) perhaps most importantly, nearly all soil taxonomists in the nineteenth through the mid-twentieth centuries have emphasized the factors of soil formation rather than the quantitative properties of soils themselves.

A. The Enormous Challenge of Mapping Soils on the 5 Billion Hectare Tropical Landscape Obtaining a representative view of the soils on 5 billion hectares of tropical lands has been an extremely difficult task. Because the origins of soil science were in the temperate zone, relatively few soil scientists continuously studied soils in the tropics. Not only is the area large, but much of the tropics is difficult to traverse, especially those humid areas covered by forests, and reports were often confined to small areas and short-term studies. To this day, the regions of the tropics that have the least reliable soil data, e.g. the Amazon and Zaire River basin, are precisely those regions which historically have been mapped with the largest area of Oxisols (Fig. 4). There have been few historical incentives for surveying soils in these areas supposed to be dominated by Oxisols. As a result, soils in the rainforests of Amazonia and in the Zaire River basin were mapped using general ideas about climate, vegetation, and physiography, i.e. by factors of soil formation, rather than by directly observed soil data. A major exception to this pattern has resulted from the Brazilian government’s decision in about 1970 to stimulate development of the Amazon basin. Systematic soil surveys (in addition to mapping and classification of vegetation, geomorphology, and land suitability) were initiated by the Brazilian government and took about a decade to complete. The results will be improved in the years to come, but in the short-term the work is nothing less than a major advance for soil science

SOIL DIVERSITY IN THE TROPICS

333

and ecology. What were large, highly uncertain mapping units of Oxisol covering most of Amazonia, have given way to a variety of soil mapping units that are based on directly observed soil profiles. The Amazonian soils of terra firme, the upland soils between the rivers, have finally been examined with some degree of comprehensiveness. These data are evaluated in detail later in this paper, because the results have much to say about soil diversity in the tropics, especially that within the humid tropics. Similar comprehensive soil surveys have yet to be conducted in the Zaire River basin of Africa.

B. Interdisciplinary Miscommunication about Soils in the Tropics At least some of the problems in developing realistic ideas about soils in the tropics are associated with poor communication among scientists of different disciplines that have a purview over natural resources of the tropics. Historically, interdisciplinary communication has been notably deficient between soil fertility specialists and soil taxonomists, and even worse between general soil scientists and ecologists. This latter deficiency is especially significant considering that soil science is fundamentally a biological and ecological science. Lack of communication between soil scientists and ecologists has allowed overgeneralizations about soils of the tropics to persist within scientific communities that have had little contact with the latest developments in the soil sciences. Large numbers, but certainly not all, of soil scientists and ecologists appear to be generally aware of the remarkable diversity of soils in the tropics. The same may not yet be the case for social scientists who work with the economics and policies that affect natural resource management in the tropics. Nevertheless, many ecologists (including soil scientists) give credence to the popular notions of “tropical soil”, although most of the same scientists would no doubt discount the meaning of a so-called “temperate soil”. Given the many years over which both natural and social sciences have largely ignored the diversity of soil resources in the tropics, an interdisciplinary effort is needed to describe and evaluate tropical soil resources more realistically. More realistic ideas about soils in the tropics are particularly appropriate for new editions of soils and terrestrial ecology textbooks.

C. The Tower of Babel Effect of Too Many Soil Taxonomies and Nomenclatures General understanding of the soils in the tropics has been stimied by the large number of languages, classification systems, and nomenclatures

FAO/UNESCO Soil Map Reliability class1

ClaSsIII (van Warnbeke et al..

Fig. 4 The reliability of 3il mapping units ,aried considerably for the FAO/UNESCO (1971) map of South America, but I marked correlation existed between Ferralsols (Oxisols) and the regions of least map reliability (Amazonia). Soil mapping units with reliability class I were based on soil surveys, in which map-unit composition and boundaries were based on actual field observations of soils. Mapping regions with reliability class I1 were based on soil reconnaissance, in which boundaries were based to a large extent on topography, geology, vegetation, and climatic data; the composition of map units were based on field observations. Reliability class 111, indicated that only general information was used to construct both boundaries and composition of each soil map unit.

Fig. 5 The FAONNESCO map of Africa relied on soil data of vastly different qualities, from detailed soil surveys such as that of Burkina Faso (reliability class I) to general information such as that of Mali (reliability class 111). See caption for Fig. 4 for description of reliability classes. As a consequence, detail of mapping unit borders and density of mapping units were much greater in Burkina Faso than in Mali.

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D. D. RICHTER AND L. I . BABBAR

that have been used by soil scientists and ecologists working in the tropics. Even by the mid-l950s, there was serious uncertainty about whether soil science was adequately developed to devise a taxonomy that was applicable not only to known types of soils in the tropics, but that was also flexible enough to accommodate new concepts and types of soils yet to be discovered (Mohr and Van Baren, 1954). In 1948, for example, an International Conference on Tropical and Subtropical Soils was held at Rothamsted Experiment Station in England, to promote mutual understanding of soils in the tropics. According to Mohr and Van Baren (1954), however, the gathering had significant communication problems caused by the numerous taxonomic and soil classification systems and technical vocabularies used by participants. Since even well informed delegates had difficulty understanding soil systems, methodologies, and terminologies that were highly parochial, the conference strongly recommended the need for a unified system of soil classification and nomenclature applicable throughout the tropics. Sanchez (1976) used the analogy of the Tower of Babel to describe the problems faced by scientists in the tropics working with soils that were described by an extremely wide variety of soil taxonomies and nomenclatures. Although there are still a large number of classification systems still in use, Soil Taxonomy and the FAO/UNESCO classification are two systems that are increasingly used, and correlation of soil types among classification systems has been an important activity for soil scientists. Table 3 contains approximate correlates of four major systems currently in use that are suited to characterize soils in the tropics: the FAO/UNESCO system, Soil Taxonomy, the Brazilian system, and the French ORSTOM classification. Technically, the FAO/UNESCO system is not a detailed taxonomic system, but rather a mapping system with a limited number of map units. Other major taxonomic systems that are used in various parts of the tropics include the Australian systems (Stevens, 1962; Northcote, 1960; Moore et al., 1983); the Belgian system of INEAC (Sys, 1960); the Soviet system (Gerisimov et al., 1974); and the 1938 USDA system (Baldwin et al., 1938; Thorp and Smith, 1949; Kellogg, 1950). Discussions about correlating soil taxa are found in Sanchez (1976), Buol et al. (1989), and Duchaufour (1982). In recent years, two systems, Soil Taxonomy and the FAO/UNESCO system have increasingly been used as the common international systems of soil classification.

D. Emphasis on Factors Rather Than on Effects of Soil Formation: The 1938 Soil Classification Despite the challenge of mapping 5 billion hectares of soils, a lack of communication among scientific disciplines, and the confusion caused by

Table 3 Taxonomic correlations among four major classification systems for soils in the tropics based on 1974 FAO/UNESCO Soil Map of the World (Sanchez, 1976; Aubert and Tarvenier, 1972; Duchaufour, 1982; Buol et al., 1989; Beinroth, 1974; Moss, 1968; Carmargo et al., 1986, Soil Survey Staff, 1987b) ~~~

1974 F A O ~ E S C OSystem

1975 (1987) Soil Taxonomy

Ferralsols

Brazilian System

French (ORSTOM) System

Oxisols

Latossolos with latosolic B < 6.5 mEq/100 g clay CECe

Sols ferralitiques fortement desatures Sffd, typiques ou humiferes

Orthic F. or Acric F.

Ustox Udox

Latossolos Vermelho Escuro

Sffd, typiques ou humiferes

Orthic F. or Acric F.

Ustox Udox

Latossolos Vermelho Amarelo

Sffd, typiques ou humiferes

Xanthix F.

Ustox Udox

Latossolos Amarelo

Sffd, typiques ou humiferes

Rhodic F.

Eutrustox or Eutrudox

Latossolos Roxo o Terra Roxa Legitima

Sffd, typiques ou humiferes derives de basalte

Acrisols

Ultisols

Podzolicos Vermelho Amarelo Distroficos

Sols ferralitiques fortement et moyennement desatures; eluvies

Lithosols

Various lithic subgroups

Solos Litolicos

Lithosols et sols lithiques

Arenosols

Psamments

Areis Quartzosas Regossolos Plintos- Sols ferralitiques moyennement on solos fortement desatures de texture sableuse

Luvisols

Alfisols

Podzolico Vermelho Amarelo Equivalente Eutrofico y Terra Roza Estruturada

Sols ferrugineux tropicaux lessives

~~~

Continued

Table 3 Continued 1974 FAO/UNESCO System

1975 (1987) Soil Taxonomy

Regosols

Brazilian System

French (ORSTOM) System

Psamments Orthents

Regossolos

Sols mineraux brut et Sols peu evolue d’apport eolien

Yermosols

Aridisols Entisols

Sands

-

Cambisols

Inceptisols

Cambissolos (incipient B horizons)

Dystric C.

Dystropepts

-

Eutric C.

Eutropepts

Sols ferralitiques faiblement desatures, rajeunis et Sols ferrugineux tropicaux

Humic C.

Humitropepts

Sols ferralitiques fortemente et moyennemnet desatures, humiferes, rajeunis

Sols ferralitiques fortement et moyennement desatures, rajeunis (p.p.)

Ultisols

Podzolicos Vermelho Amarelo

Sols ferralitiques fortemente et moyennement desatures; eluvies

Alfisols

Podzolico Vermelho Amarelo equivalente y Terra Roza Estruturada

Sols ferrugineux tropicaux lessives

Vertisols

Vertisols

Vertissolos (Grumusols)

Vertisols

Gleysols

Various aquic suborders

Gleissolos

Sols hydromorphes

Nitosols

Xerosols

Mollic Aridisols

Sands

Fluvisols

Fluvents

Solos Aluviais

Sols mineraux brut et Sols peu evolue d’apport alluvial et colluvial

Planosols

Paleudalfs Paleustalfs Aqualfs Aquults Argids Argialbolls

Planossolos (hardpan below A horizon)

Sols ferrugineux tropicaux lessives (P.P.)

Kastanozems

Ustolls

Andosols

Andisols

Histosols

Histosols

Solos Organicos

Sols hydromorphes

Solonchaks

Aridisols

Solos Salinos (natric B horizon)

Sols halomorphes

Rendzinas

Rendolls

-

Sols calcimagnes imorphiques

Phaeozems

Udolls Aquolls

Solonetzes

Natargids Natrustalfs Natralbolls

Solonetz Solodizados (natric B horizons)

Sols halomorphes

Podzols

Spodosols

Podzols

Podzols

-

Andosols

Sols isohumiques

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D. D. RICHTER A N D L. I . B A B B A R

too many soil taxonomies, perhaps the single greatest impediment to the development of realistic ideas about soils in the tropics was the general direction in which soil science itself developed as a science. In the mid and late nineteenth century, Dokuchaev in Russia and Hilgard in the USA originated the seminal ideas about how the complex of environmental factors controls soil development. Both, however, actually stressed climatic factors in soil formation, an emphasis that diminished the importance of factors such as geology, topography, and soil age (Hilgard, 1892; Sibirtzev, 1914). The ideas of Dokuchaev and Hilgard dominated the developing science of soils and they helped create the momentum to organize early soil taxonomies according to environmental factors of soil formation (Basinski, 1959). This approach proved more a taxonomy of the factors that formed soils, than a taxonomy of soils themselves. Fundamental classification problems were created by basing soil taxonomies on soil-forming factors rather than on measurable soil properties. Worst of all, few if any quantitative criteria were set for soil taxa, and even experts had difficulties in classifying soils consistently. By 1900, Sibirtzev (1914) enlarged the climatic emphasis of Dokuchaev to emphasize effects of both biology and climate on soil formation. As a part of this perspective, Sibirtzev (1914) promoted an idea of “zonal soils”, a concept that came to dominate soil and ecological science for much of the late-nineteenth and twentieth centuries (Marbut, 1935; Whittaker, 1975). “Zonal soils” were actually an extension of Dokuchaev’s classification of “normal soils”: well drained, upland soils with well-developed horizons, that were supposed to reflect regional climatic regimes. These ideas are illustrated in Fig. 1. Major objections were raised by Milne (1935a,b) to the zonal and normal soil concepts. Milne (1935a,b) described repeating soil catenas in west Africa, and questioned whether true “zonal” soils were those on the uplands or those in the swales, because soils of both landscape positions were in a dynamic balance with regional and local climates. The concepts of zonal and normal soils demonstrated that bioclimatic factors were clearly overemphasized at the expense of the entire complex of interactive environmental factors that control soil formation. Perhaps it is instructive that Marbut’s (1935) map of “non-normal” soils of the USA (not including Alaska) included about half of the USA. Included as “non-normal” soils were poorly drained soils; soils with imperfectly developed profiles; unstable soils in mountainous landscapes; soils with fragipans, claypans, or any indurated hardpan horizons; and soils with high limestone contents. Effects of geologic materials, landforms, and hydrology were especially diminished in importance by the zonal concept. The priority that the zonal concept gave to upland soils may also have contributed to

SOIL DIVERSITY IN THE TROPICS

34 1

other soil conceptual problems, such as those mentioned by Daniels and Nelson (1987): that soil variability influenced by stratigraphy, geomorphology, and hydrology has yet to be adequately appreciated, even by most soil scientists in the late twentieth century. Some soils and ecological scientists still employ zonal concepts (e.g. Whittaker, 1975; Burnham, 1984), but from most modern perspectives of soils, concepts of zonal and normal soils are too arbitrary to be useful. On balance it even might be argued that the zonal and normal soil concepts are intellectual constructs that have obscured as much as they have enlightened. The ideas of Sibirtzev (1914) and Marbut (1927, 1935) about zonal and normal soils had a great influence on the 1938 USDA soil classification, the taxonomic system that has at least indirectly promoted the popular notion of “tropical soil”. The authors of the 1938 USDA system, led by C. E. Kellogg, were given only a short time by the US Secretary of Agriculture, Henry Wallace, to prepare a comprehensive classification of soils to be published in the 1938 Agricultural Yearbook, Soils and Men (Baldwin et a l . , 1938; Forbes, 1986). Kellogg’s group relied heavily on the ideas of Marbut (1927, 1935), and given time constraints of publication, the group developed few new concepts of soil classification. Although the 1938 USDA system was later refined (Thorp and Smith, 1949; Rieken and Smith, 1949; Kellogg, 1950), the system in many ways solidified past ideas about soils. The 1938 system and its revisions were entirely within the old paradigm of soil taxonomy in which soils were classified based on inferred factors of soil formation. The 1938 system had three major categories that were operationally useful: soil orders, soil great groups, and soil series. There were three soil orders: zonal, azonal, and intrazonal soils, categories that were taken directly from Sirbitzev (1914) and can be traced back to ideas of Dokuchaev. There were up to 36 great soil groups, up to 20 of which were zonal soils, some of which are illustrated in Fig. 1. Each great group contained soils with broadly similar soil profiles that reflected generally similar formation processes. The number of soil series are of course indefinite. With use, many shortcomings were found with the 1938 system. A lack of specific and quantitative limits for each great group allowed different experts to classify the same soil series in different great groups. Moreover, the system was not able to accommodate changes in conceptual ideas very easily. Despite important revisions (Thorp and Smith, 1949; Rieken and Smith, 1949; Kellogg, 1950), the system simply had difficulty in being applied or in being updated effectively (Beinroth, 1974). The 1938 system was nonetheless very widely publicized in the USA

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D. D. RICHTER A N D L. I . BABBAR

and internationally, and was promulgated in major textbooks of soil science, ecology, agronomy, biology, forestry, as well as other disciplines. As a result, the 1938 system has promoted, especially among non-soil scientists, a relatively monolithic concept of zonal tropical soils, the red laterite of Fig. 1. For example, Fig. 1 is from the classic ecology textbook,The Study of Plant Communities by Oosting (1956), which like other texts emphasized the effects of regional climates on vegetation types and on zonal soils. Similar diagrams were rationalized by Lutz and Chandler (1946) in Forest Soils, by Whittaker (1975) in a contemporary plant ecology text, Communities and Ecosystems, and in Odum (1971) in Fundamentals of Ecology. Several generations of students have thus been educated entirely within the 1938 system. The system has had a profound effect on the way natural and social scientists think about soils, and specifically about soils in the tropics. In fact, many concepts and great soil groups of the 1938 system, such as the Latosol, continue to be used with little modification. Whittaker (1975) justified the use of the 1938 system in his classic text Communities and Ecosystems because as he argued, useful ideas about “typical” soil and ecosystem conditions could be communicated easily by using great soil groups that were associated with “typical” ecosystems. Whittaker (1975) described the tropical rainforest type as a forest supported by Latosol, a forest with a “rich nutrient economy perched (italics added) on a nutrient-poor substrate.” While important rainforest ecosystems are supported by nutrient-poor Latosols (i.e. Oxisols), Whittaker’s (1975) typification approach can easily diminish the importance of the diversity of soil taxa and edaphic conditions that support tropical rainforests, and the enormous areas of tropical rainforest supported by non-Latosol (non-Oxisols) soils (Whitmore, 1984; Cochran et al., 1985). Indirectly, the typification approach has narrowed our perspective of soil and forest vegetation in the tropics, a result that may have major practical consequences. Moran (1981), for example, argued that a narrow perspective of tropical rainforest soils has caused the potential for agricultural and forestry development to be greatly underestimated.

IV. ADVANCES IN SOIL TAXONOMY AND THE CREATION OF THE WORLD SOIL MAP A. Soil Taxonomy: A new Scientific Paradigm To understand modern perspectives of soil diversity, one needs at least some understanding of the structure and development of modern soil taxonomy and classification. By the 1950s, an international group of soil

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taxonomists initiated a new soil classification system that radically altered the taxonomy of soil. Development of the new system continued through the 1960s and 1970s. In 1975, the definitive proposal for the new system was published as Soil Taxonomy (Soil Survey Staff, 1975). Major amendments to Soil Taxonomy have been published since 1975 (Soil Survey Staff, 1987a), amendments which have demonstrated the vitality and evolution of this new approach and system. Although some soil concepts in the Soil Taxonomy system are inherited from the 1938 USDA system (Smith, 1983), the new system breaks from the past because it is quantitatively based. Prior to Soil Taxonomy, soil forming factors were emphasized in taxonomic systems; however, following the development of Soil Taxonomy, the effects of soil forming factors, i.e. quantifiable soil properties, are emphasized by most soil taxonomies. Currently, the FAO/UNESCO, Brazilian, and Australian systems are all quantitatively based, in contrast to the Soviet and French (ORSTOM) soil classification systems which are still heavily influenced by soil formation ideas (Buol et al., 1989). The Soil Taxonomy system is a flexible, multicategorical system, that is based on key soil horizons, i.e. diagnostic soil horizons, characterized by their physical, morphological, and chemical properties. Diagnostic soil horizons include mollic, histic, oxic, argillic, spodic and kandic horizons each of which is briefly explained in Table 4. In Soil Taxonomy, diagnostic soil horizons help distinguish many soils at the highest taxonomic level, i.e., the order level, as is illustrated in Table 4. A number of secondary diagnostic horizons of surface and subsoils are used to differentiate groups of soils below the order level. Twenty-one of 22 diagnostic soil horizons listed in Soil Taxonomy (Soil Survey Staff, 1975) are found in soils of the tropics, an indication of the diversity of soil taxa found in the tropics. One of the outstanding features of the Soil Taxonomy system is its conscious openendedness; a system that intentionally accommodates change and improvement without drastic revision of the structure of nomenclature. In contrast, the structure of the 1938 USDA system was typical of many older soil taxonomy systems, in that it could not survive pressures created by the expanding scientific knowledge about soils. A seminal and continuing idea of the Soil Taxonomy system was to devise a structure that would not self-destruct as new information was assimilated. Although Soil Taxonomy is not complete, it is not meant to be. A major intention of its publication was to have it used and criticized by as many people as possible (Smith, 1983). The system evolves to meet the wide range of soil conditions found throughout the world. The Soil Taxonomy system is highly organized. Its nomenclature appears complex initially, but once the system is even partly understood,

Table 4 Primary diagnostic soil horizons in Soil Taxonomy used to classify soils in both Soil Taxonomy (Soil Survey Staff, 1986, 1987a) and the FAO/UNESCO (1974) system.

Soil Taxonomy Diagnostic Horizon

Location within Profile

Soil Taxonomy Order

1974 FAO/UNESCO Primary Map Unit

Brief Description

Mollic

Surface

Mollisol

Kastanozem Phaeozem Rendzina

Dark, deep (> 25 cm), friable, organic rich, with relatively high pH, and nutrient content.

Histic

Surface

Histosol

Histosol

Deep, accumulation of organic matter. Peat or muck.

Oxic

Subsurface

Oxisol

Ferralsol

Low cation exchange capacity with Fe and Al-oxide rich, variable charge, low activity clays; highly structured friable.

Argdlac

Subsurface

Alfisol Ultisol

Luvisol Acrisol Nitosol

Illuvial clay, with high or low acidity (Ultisol or Alfisol, respectively). Nitosols have high argillic clays in subsoils.

Spodic

Subsurface

Spodosol

Podzol

Acidic coarse textured horizon, high in organic matter and amorphous Fe and Al oxides.

Kandic

Subsurface

Alfisol Ultisol

-

Mutually exclusive from oxic but not of argillic horizons. Sharp increase in low activity clay with depth; no argillans evident.

SOIL DIVERSITY IN THE TROPICS

345

much of the nomenclature can be translated, if the person has some experience with soils. It begins with eleven upper-level classes, called orders (Table l ) , and the remaining levels of the hierarchy are suborders, great groups, subgroups, families, and series. Specifying a soil at any level above the family automatically specifies the higher hierarchial classes to which that soil belongs. This nomenclature has the appearance of complexity, but actually contributes greatly to communication. All subdivisions of a soil order end in a characteristic syllable: for Oxisols this is -ox, for Ultisols this is -ult. Buol et al. (1989) give introductions to the different levels of the system, and more complete descriptions of the system are found in Soil Taxonomy and its amendments (Soil Survey Staff, 1975, 1986, 1987a,b). In the years since the publication of Soil Taxonomy, the system has received much use throughout the tropical world. The amendments published since 1975 are most significant because many refine concepts of soils found in the humid subtropics and tropics. Many of the amendments have originated from scientists based in the tropical world (Moormann, 1985; Forbes, 1986). A new diagnostic horizon, the kandic horizon, has been recently accepted for Ultisol, Alfisol, and Oxisol orders (Soil Survey Staff, 1986, 1987a). Major revisions have been made in the Oxisol order (Buol and Eswaran, 1988), and a new soil order, the Andisols, has been adopted to include many soils of volcanic ash origin (Leamy, 1988). Each of these three changes (the kandic horizon, Oxisol revisions, and the new Andisol order) have made the Soil Taxonomy system significantly more applicable to soils in the tropics and each are briefly described here. The kandic subsoil horizon is a new diagnostic horizon, recently introduced to ease problems in differentiating Oxisols, Ultisols, and Alfisols that have subsoils with low activity kaolinitic clay. The kandic horizon shares properties with both argillic and the oxic horizons; it is mutually exclusive with oxic horizons, but not with argillic horizons. Kandic horizons have sharp increases in clays between surface and subsoils (similar to argillic horizons required for Alfisols and Ultisols), and need not have clay skins (peds that are coated with clays) that indicate active clay movement from surface to subsoils. Kandic horizons must have clays of very low buffering capacities, i.e., C 160 mmol,/kg of clay of CEC determined at pH 7. The 1987 revisions of Soil Taxonomy make the definition of Oxisol much more detailed and have slightly broadened the Oxisol concept compared with that of 1975 (Soil Survey Staff, 1975, 1987b; Buol and Eswaran, 1988). Changes in concepts have been based on the wealth of data and experience with soil in the tropics that have gained in the last decades. Such revisions have caused some confusion; e.g. Van Wam-

346

D. D. RICHTER AND L. I . EAEEAR

beke (1989) questioned why some high clay soils (> 40% clay) with clayey accumulations in subsoils are now included as Oxisols. For the most part, however, the 1987 Oxisol criteria (Soil Survey Staff, 1987b) combined with the new kandic horizon should make classification of low activity clay soils more straightforward. A recent significant change in the Oxisol order was the replacement of Udox for the older Orthox (“orth” meaning the “true” or the “common” Oxisol). Buol and Eswaran (1988) describe how this change responded to the concern that more Oxisols may actually be located in drier climate regimes than previously appreciated: the so-called ortho-Oxisol may well be the Ust-ox (an Oxisol soil moist enough to grow crops for 90-270 days/year) rather than the Ud-ox (an Oxisol soil moist enough to grow crops for > 270 days/year). The new 1987 revisions of Oxisols have also added considerable detail to the order. In 1975 there were 30 subgroup taxa (the fourth level of the Soil Taxonomy system hierarchy), whereas according to Buol and Eswaran (1988), the 1987 Oxisol order has 212 subgroups. Evidence of the flexibility of the Soil Taxonomy system is the recent inclusion of an 11th soil order, the Andisols or volcanic-ash soils. This is the first additional soils order added to the system since publication of Soil Taxonomy (Soil Survey Staff, 1975). Soil and ecological scientists throughout the world, many from the tropics, have argued that chemical and physical properties of such volcanic soils (formerly classified as Andepts, a suborder of the Inceptisol order), warranted an upgrade to order status to be adequately classified (Forbes, 1986; Smith, 1968).

B. FAO/UNESCO Soil Map of the World Because of the practical and theoretical importance of soil taxonomy, the classification and mapping of the world’s soils have a long-term activity of the International Society of Soil Science (ISSS). Marbut (1927) presented an initial scheme for world soil classification at the 1st International Congress of ISSS. Mapping of the world’s soils emerged as a top priority of the 6th International Congress held in Paris in 1956. At the 7th International Congress held in Madison, Wisconsin in 1960, soil maps of each continents were presented at scales of 15 million to 1:lO million. At the Madison meeting it was recognized that the complexity of the task had been previously underestimated. Nomenclature, survey methods, legends, and systems of classifications varied so widely that comparisons between and within continents were very limited in their usefulness. In response to recommendations of the 7th ISSS Congress, F A 0 and

SOIL DIVERSITY IN THE TROPICS

347

UNESCO agreed to join ISSS in preparing a comprehensive world soil map to be based on the most complete soil survey data and field correlation available. The FAO/UNESCO world maps were eventually based on over 600 individual soil maps and many different mapping systems. Considerable attention had to be given to development of the map legend, mapping units, and nomenclature (FAO/UNESCO, 1974). It is no coincidence that like the Soil Taxonomy system, FAO/UNESCO categories were based not on speculative concepts of soil information, but on quantitative criteria of diagnostic soil horizons (FAO/UNESCO, 1974). This was clear evidence for the success of the new quantitative soil taxonomic paradigm. To quantify soil diversity in the tropics we will evaluate mapping data of the FAO/UNESCO map of the world. Although the FAO/UNESCO map unquestionably contains the most comprehensive soil data of the tropics, the classification system has three characteristics that are important to interpreting the data of the FAO/UNESCO world soil map: (1) the system is not a complete classification system; it has at most three hierarchical levels; (2) the system is not very flexible, although the original legend, but not the map, has been revised once (FAO/ UNESCO, 1988); and (3) the system is based on national and regional soil surveys that have a wide range of quality.

1. A Limited Soil Classification The FAO/UNESCO system is not a complete soil classification; it has only two and at most three levels of organization. Primary FAO/ UNESCO map units like the soil orders of the Soil Taxonomy system, are grouped by diagnostic soil horizons. For example, the primary map unit of Ferralsol includes soils with a diagnostic oxic-like ferralic horizon, whereas the primary map unit of Acrisols and Luvisols (mainly Ultisols and Alfisols, respectively) includes soils with diagnostic clay-enriched (argillic) subsoils that are acid or non-acid, respectively. Secondary map units of the FAO/UNESCO system are used to account for variation within primary units. Humic Ferralsols o r Humic Acrisols are Ferralsols or Acrisols with deep surface soils (> 25 cm) and large carbon accumulation throughout the surface soil (> 0.6% carbon throughout the surface horizon). Soil texture and slope data are also included in mapping units. The FAO/UNESCO system, however, as a two and at most three-tiered system, contains no application to local soil series. At 1:5 million or 1:l million scale, it has no ability to serve local land-use purposes. The system should not be criticized for objectives it does not claim to have, for it supplies considerable information for a two- or three-level classification system used to map soils.

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D . D . RICHTER A N D L. I . BABBAR

2. System Flexibility The FAO/UNESCO system lacks much of the flexibility possessed by the Soil Taxonomy system. The FAO/UNESCO system has been updated once since its publication in 1974 (FAO/UNESCO, 1988), although the maps remain as originally published. The next soil map of the world may not be completed until the twenty-first century, due mainly to the magnitude and expense of a world soil mapping project. Nonetheless, despite being relatively inflexible, the FAO/UNESCO world map contains the best current summary of global-scale soil taxonomic data. It is, however, important to appreciate its limitations, most important of which is that the quality of the mapping varies greatly between regions, especially in the tropics.

3. Quality of Data Base A major limitation of the FAO/UNESCO map is that the map is based on soil data of a wide range of primary sources, ranging from systematic soil surveys to general surveys which were based on practically no soil data at all. Most of the mapping units of the soils map of the world were entirely dependent on soil mapping that was accomplished prior to about 1975. Figure 5 is taken from the FAO/UNESCO map of west Africa, a region that has soils mapped with relatively detailed resolution in Burkina Faso and with much less resolution in Mali. The figure illustrates how differences in map resolution of the original soils surveys used in the FAO/UNESCO project had a direct effect on the size and detail of mapping units used in this soil map of Africa. Primary soil surveys of Burkina Faso were relatively detailed compared with those available for Mali. Apparent soil taxonomic diversity shown on the soil map of the world can be due to political boundaries between nations rather than real soil differences (Fig. 5 ) . The recent revisions of the FAO/UNESCO system (FAODNESCO, 1988), are important for understanding soils in the tropics and for anticipating the future direction of soil taxonomic concepts. The 1974 FAO/UNESCO system was similar to the 1975 Soil Taxonomy in many ways and some of the recent revisions made in Soil Taxonomy (Soil Survey Staff, 1986, 1987a, 1987b) were similar to those recent revisions made in the FAO/UNESCO (1988) system. Significantly, recent changes in the FAO/UNESCO and Soil Taxonomy system also suggest that they are moving away from each other, since similar classification problems were solved in markedly different ways. Two diagnostic horizons that have often been difficult to distinguish in the field, the oxic and the argillic, were revised and even renamed by the revised FAO/UNESCO

SOIL DIVERSITY IN THE TROPICS

349

(1988) system, the ferralic and the argic, respectively. The new ferralic B horizon now contains all soils with very low activity clays, regardless of textural differences between surface and subsoil. The ferralic B also contains very low concentration of weatherable minerals, low water-dispersible clays, low silt-clay ratios, and < 5% by volume that shows any rock structure. The argic B is now subordinate to the ferralic, and represents most fine-textured subsoils that underlie relatively coarse-textured surface soils. In contrast, the Soil Taxonomy system has maintained both oxic and argillic concepts, but added the kandic concept in an effort to distinguish between low activity argillic from the oxic horizon (Table 4). A second revision that moves the FAO/UNESCO system further away from the Soil Taxonomy system, is the elimination of nearly all criteria based on transient soil properties that respond to climate. Although the 1974 FAO/UNESCO system contained little reference to soil moisture or temperature, such climatic regimes are central to the Soil Taxonomy system’s practical objectives of soil and ecosystem management. At a very high level (the suborder), the two systems have no direct correlates. The French and Brazilian systems of classification also include criteria that only involve relatively permanent soil morphologic, chemical, or physical properties, i.e., no soil moisture or temperature-regime data are included. Taxonomically, such an approach by the FAO/UNESCO system is extreme, in that it seems to overestimate the permanence of soil properties as well as the difficulty of distinguishing between soil moisture and temperature regimes. Van Wambeke (1989) severely criticizes this movement away from using soil dynamic properties (such as soil moisture regime) to classify soils, due to the ecological importance of such soil properties. In summary, the advancement of soil taxonomy has been a notably difficult task, especially in developing a framework for soils of the tropics. The progress in soil taxonomy is, however, very impressive, given that soils are difficult to observe and highly variable and local in the expression of their properties. The development of the Soil Taxonomy system and the completion of the FAO/UNESCO Soil Map of the World have both contributed to our understanding of soil diversity in the tropics. Although the systems are not suited for all soil survey applications (nor should they be expected to), both systems are recognized increasingly as systems that can classify soils repeatably throughout the world and in ways that have meaning to land management and ecology (Isbell, 1984). It appears, however, strongly desirable that future revisions of Soil Taxonomy and the FAO/UNESCO classifications be much more co-ordinated, especially concerning issues that bear on the soils of the tropics.

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D. D . RICHTER AND L. I. EABBAR

V. DIVERSITY OF SOIL TAXA IN THE TROPICS At a global scale, soil taxa in the tropics are diverse. Even on soil maps of the tropics with 1 5 0 million scale, all ten of the original ten soil orders of the Soil Taxonomy system are mappable (Aubert and Tavernier, 1972). At this scale, each of eight soil orders exceed 50 million hectares of the tropics (Aubert and Tavernier (1972); Drosdoff as cited by Sanchez (1976)). According to such maps, common soils in the tropics (Table 5) include soils that are iron and aluminium-rich, friable, and well drained (Oxisols); soils that are sandy, droughty, and occasionally salt-rich (Aridisols) ; soils that have clay-enriched subsoils that are either nutrient-poor or nutrient-rich (Ultisols or Alfisols, respectively); soils with high-water tables (Aquepts, Aquults, and Aquents); rocky soils on unstable slopes (Inceptisols or Entisols); sandy soils deposited by rivers or on coastal plains (Psamments); clayey soils that are nutrient-rich and have clays that shrink and swell upon drying and wetting (Vertisols); deep, organic-rich soils that are often highly fertile (Mollisols); and soils of deep accumulations of peat (Histisols). To quantify the diversity of soil types found on a regional scale throughout tropical Africa, America and Asia, we summarized areal estimates of all soil map units found in each of the five volumes of the world soil map containing tropical areas, Volumes 111, IV, VI, VII, and IX (FAO/UNESCO, 1971, 1975, 1977a,b, 1979). Our approach was to collate soil-area data that are found in each of the five volumes of the 1974 FAO/UNESCO system, based on nations that are mainly of lower latitudes than the Tropics of Cancer or Capricorn. Tropical America included all nations in Central and South America including Mexico, but excluding Argentina, Chile, and Uruguay. Tropical Africa included all nations except South Africa, Morocco, Algeria, Libya, and Egypt. Tropical Asia included India and all nations of the Southeast Asia peninsula and islands including Indonesia, the Philippines, and New Guinea. Our rationale was to quantify the very wide range of soil types found throughout tropical Africa, America and Asia, and our results are presented in Table 6 and 7 for primary and secondary categorical levels, respectively. All descriptions and analyses in this section of the paper (Part V) use the 1974 FAO/UNESCO legend except where stated. We use the world soil map data ( 1 5 million scale) to demonstrate diversity of soil taxa in the tropics as a whole. The FAO/UNESCO map is not adequate in much of the humid tropics, and recent Brazilian soil survey data of the Amazon River basin (at 1 5 miilion scale) are analysed later in this paper to demonstrate that soils in the tropics are

35 1

SOIL DIVERSITY IN THE TROPICS

Table 5 Areas of major soils in the tropics as estimated at 150 million scale by Drosdoff, Aubert and Tavernier (after Sanchez, 1976).

Soil Order or Suborder Aridisols Ustalfs

Orthoxs Mountain soil complexes Udults Psamments Ustoxs Aquepts Tropepts Ustults Usterts Mollisols Aquults Udalfs Aquents

TOTAL

Description Desert soil, dry > 50% of most years Clayey, nutrient-rich, nonacid subsoil; dry > 90 days, < 180 dayslyear Extr. weathered; no clayey B hor; moist > 270 days/year Various; rocky, shallow, steep, volcanic ash. Clayey, nutrient-poor, acid sub-soil > 270 dayslyear Deep sand, well drained, weakly developed horizons Extr. weathered; no clayey B hor; dry > 90 days, < 180 dayslyear Weakly developed, young soil; water table at surface each year Weakly developed, young soil; isotropical temperature regime Clayey, nutrient-poor, acid subsoil; dry > 90 days, < 180 dayslyear Shrink-swell clays, neutral pH; dry > 90 days, < 180 days /year Deep organic-rich soils of grasslands Clayey, nutrient-poor, acid subsoil; water table at surface each year Clayey, nutrient-rich, nonacid subsoil; moist > 270 dayslyear Weakly developed, young soil; water table at surface each year

Area millions of ha

Percent of Tropics

900

18.4

760

15.4

750

15.0

600

12.2

410

8.2

390

8.0

350

7.5

285

6.0

115

2.3

100

2.2

100

2.0

50

1.o

40

1.o

40

0.8

10

0.2

4900

100.2

Table 6 Areas of major soil mapping units of the l:5-million sacle FAO/UNESCO Soil Map of the World for nations mainly within the Tropics of Capricorn and Cancer.

Primary Map Unit

Short Description

Central America

Africa

South America

Asia

Total

Percent of total

loooS of hectares

Ferralsol

Sesquioxide-rich clay (Oxisols)

Acrisol

650

417 640

614 520

15 120

1047930

21.2

Acid, argillic (Ultisols)

21 310

84 960

189600

250 590

546460

11.1

Lithosol

Rocky, shallow (Lithic subgroups)

24 660

234 080

145520

83 OOO

487 260

9.9

Arenosol

Sand (Psamments)

0

371 340

72 395

36 600

480 340

9.7

Luvisol

Non-acid, argdlic (Alfisols)

35 460

218 860

118060

100280

472 670

9.6

Regosol

Thin soil over unconsolidated matter (Orthents)

13550

159470

154990

7 873

335 880

6.8

Yermosol Cambisol

Desert soils (Aridisols)

24 640

217 060

2 420

28 300

272 420

5.5

Incipient change in structure, consistence (Inceptisol)

30 810

100770

24 190

87 620

243 390

4.9

Nitosol

Low CEC in argilhc (Alfisol, Ultisol)

10790

115390

20210

45 360

191750

3.9

Vertisol

Shrink-swell, clayey (Vertisol)

16311

89 810

7 680

66840

180640

3.7 2.9

Gleysol

Reduced horizons due to wetness (Aquepts, Aquents)

Xerosol

Dry soils of semi-arid regions (Aridisols) Alluvial soils (Aquents)

Fluvisol

6 350

47 600

57 440

32 250

143640

14 980

105760

14340

0

135080

2.7

3 140

57 680

24 490

48 290

133590

2.7

Planosol

Poorly drained with abrupt A-B boundary (Alfisols, Ultisols)

1520

8360

40560

11430

61870

1.2

Kastanozem

Organic-rich with low acidity (Mollisols) 36 OOO

20

17 810

0

53840

1.1

Andosol

Volcanic-ash, high organic, amorphous, 19 530 (Andisols)

4 820

11320

8880

44540

0.9

Histosol

Organic soils (Histosols)

2 490

1750

2420

24570

31230

0.6

Solonchak

High soluble salt concentrations (Aridisols)

240

10940

6 700

4900

22780

0.5

Shallow soil over limestone (Udolls and 13 550

589

0

3440

17 580

0.4

Rendzina

Aquolls)

Phaeozem

Thick, base-rich organic horizon (Mollisol)

1090

340

11010

1780

14220

0.3

Solonetz

High sodium salt concentrations (Aridisols)

0

9 360

2 970

0

12 340

0.2

Podzol

Light E horizon, subsoil accumulation of Al, Fe, and organic matter (Spodosols)

0

2 320

0

3 500

5 820

0.1

277 080

2 258 920

1538 650

860 630

4 935 270

Total

-

Table 7 Areas of secondary soil mapping units in 1000s of hectares of the 1:s million scale FAO/UNESCO soil map of the world for nations mainly between the Tropics of Capricorn and Cancer

Primary Mapping Unit

Secondary Unit

Ferralsol

Acric Humic Orthic Plinthic Rhodic Xanthic Ferric Gleyic Humic Orthic Plinthic

Acrisol

Lithosol Arenosol

Central America 499 152

523 342 8269 10 937 1236 24 660

Luvic Albic Cambic Ferralic

Luvisol Chromic Ferric Gleyic Calcic Orthic Plinthic Vertic

16 927 3340 1441

Calcaric Dystric Eutric

3743 545 9266 20 344 2407 1885

13 182

Cambisol

Nitosol

Haplic Calcic Gypsic Chromic Dystric Eutric Ferralic Gleyic Humic Calcic Vertic D ystric Eutric Humic

8 160 6786 8881

4647 2336 4610 6181

Vertisol Chromic Pellic

3278 236 096 35 123 33 338 109 804 58 929

South America

Asia

Total

68433 15647 229617

221 5852 4416

23257 277565

3464 1167 32 188 16317 14 944 180 772 6366 83003

69 153 24 777 470 28 1 35 123 60 059 388 536 91 640 17 113 23 213 334 006 80 484 487 255 81 643 9679 223 416 165 598 88 1 130 895 253 102 20 542 4409 40 878 20 103 1863 1929 41 061 27 596 265 295 157 404 62 353 28 584 24 083 41 063 38 139 73 185 22 609 3521 30 811 17 883 16 179 108 480 71 327 11943 1706 105 988 72 9.50

454 20 062 5966 234 075 81 643 5163 190 347 94 185 88 1 18.517 162643 17 758 4325 207 14533

122 235 66916 145517 1422 70973 41922 65479 5249 5409

575

Regosol

Yermosol

Africa

58 16 253

1929 35 921 20 150 101466 136 676 40 771 15531 24 083 31 378 6288 23 584 17 076 105 7591 9479 5265 63 370 40 438 11581 1706 52 339 35 715

2792 152196 384 2038

11320 1383 9120 2369 12868 7343 2934 4749

3094 33 069 440 53529 21640 1343 84 22240 161 1288 1397 4109 236717137 11168 1525 13745 39337 5533 3416 14100 1388 8578 27632 17365 362 50607 16233

Continued

Table 7 Continued

Primary Mapping Unit

Secondary Unit

Central America

Calcaric Dystric Eutric Humic Mollic Plinthic

339 330 2019

G1eyso1

2852 807

Haplic Calcic Luvic Gypsic

750 6840 7385

Calcaric D ystric Eutric Thionic

354 69 1 2092

Dystric Eutric Humic Calcic Mollic Solodic

349 628 218 330

Haplic Calcic Luvic Humic Mollic Ochric Vitric

12 075 5320 18 610 1624 1203 898 15 807

Dystric Eutric

2492

Fluvisol

Planosol

Kastanozem

Solonchak Gleyic Orthic Takyric

Solonetz Podzol Total

Asia

Luvic Haplic Mollic Orthic Gleyic Humic

242 13 550 460 626

2319

1400 2104

2 258 920

1538 650 860 630

4 935 270

10713 23 738 3302 1646 1970 42 980 35 732 999 24 083 4036 11059 2624 34 864 5099 82 1 12 457 1

47 763 4476 742 4461

2578 4242 25 433

10 432 3454 457

4538 19 768 183

10813 15 092 15 981 6399

1051 31 033 6987 4446

295 1 22

158 2824 1695 144 508 1237 1708 740 8285 212 589

8236 237 10 196 55 7560 4364 1762 5189 2421

3066 898 3442 1470 17518 7050

6288 408

4163 733 3444

11012

1775

343 9363

277 080

Total

8204 2917 63 048 55 666 742 10 615 2453 1970 54 162 46 026 8841 24083 4036 22 226 22 945 72 705 11681 82 1 1421 36 232 7205 4446 8236 3518 22 22 271 5375 26 170 9212 6687 6035 22 610 2421 20518 8287 1708 11191 9668 212 17 583 11472 2744 820 11515 1400 4423

Histosol

Rendzina Phaeozem

South America

8204

Xerosol

Andosol

Africa

820 2152

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D . D . RICHTER AND L. I . BABBAR

even more diverse than that indicated by the 1974 FAO/UNESCO world map. For these latter analyses, a geographic information system (GIS) was used to estimate areal extents of various mapping units in the 500 million hectare Brazilian Amazon by comparing the older FAO/ UNESCO soil map with that of the recent Brazilian efforts (EMBRAPA, 1981; Carmargo et al., 1986).

A. General Soil Taxonomic Variation in Tropical Africa, America, and Asia Soil taxa found in tropical America, Asia, and Africa indicate wide variation of soil conditions among continents (Table 6). Tropical Asia is mapped by the FAO/UNESCO (1977b, 1979) mainly as acidic Acrisols; nutrient-rich Luvisols; weakly developed Camhisols; rocky shallow Lithosols; and shrink-swell prone Vertisols (Table 6). Most of tropical Africa is mapped by the FAO/UNESCO (1977a) as extremely weathered Ferralsols; sandy, droughty Arenosols; Lithosols; weakly developed desert Yermosols; and Luvisols (Table 6). Most of tropical South America is mapped by FAO/UNESCO (1971) as extremely weathered Ferralsols; acidic Acrisols; weakly developed and sedimentary Regosols; and rocky Lithosols (Table 6). Central America is mapped mainly as organic-rich Kastanoszems; nutrient-rich Luvisols; acidic Acrisols; weakly developed Cambisols; rocky, shallow Lithosols; and volcanic-ash Andosols. The Soil Taxonomy system correlates of these FAO/UNESCO mapping units are for tropical Asia mainly Ultisols, Alfisols, Inceptisols, lithic soils, and Vertisols; for tropical Africa mainly Oxisols, Entisols, Aridisols, and Alfisols; for tropical South America mainly Oxisols, Ultisols, Entisols; and lithic soils; and for Central America mainly Mollisols, Alfisols, Ultisols, Inceptisols, lithic soils, and Andisols. The FAO/UNESCO maps of soils in the tropics include a total of 22 primary mapping units and 97 secondary mapping units. In tropical Africa, South America, and Asia, and Central America, the total number of primary FAO/UNESCO mapping units total 22, 20, 19, and 19, respectively. Secondary mapping units total 81, 60, 69, and 57, respectively. Although large numbers of taxonomic classes are not necessarily equivalent to wide soil diversity, the specific collection of soil taxa demonstrate a marked diversity of soil between and within continental areas, considering the marked differences among many of these soils and the 1:5 million scale of these soil maps. Many of these mapping units also contain substantial within taxa variation in soil properties. The following discussions are summaries of major soil mapping units

SOIL DIVERSITY IN THE TROPICS

357

of tropics according to our GIS summary and analysis of FAO/ UNESCO (1971, 1975, 1977a,b, 1979) primary and secondary map-unit data. The summaries include definitions of each primary mapping unit, a description of the variation in soil properties within each primary mapping unit, and an estimate of the areal and geographic distribution of these soils. Tables 6 and 7 include areal estimates of primary and secondary units arranged from largest ot smallest in areal extent according to the FAO/UNESCO world map.

B. Ferralsols: Modern Inheritors of the Latosols Concept Ferralsols of the FAO/UNESCO system are generally the most weathered of all soils; they are similar to Oxisols in the Soil Taxonomy system. Ferralsol profiles are often exceptionally deep, as was well documented by Eswaran and Wong (1978) who separated the top 20 m of a weathered profile into four layers above bedrock granite in a Malaysian Udox (an Oxisol with a humid climate regime). Ferralsols are often well aggregated, porous, and friable. The diagnostic ferralic horizon, which results from very strong weathering, has low CEC that is often an important consideration for management. These horizons are dominated by sesquioxides of aluminium and iron, and often accumulate other sparingly soluable metal oxides, for example, of titanium, and more occasionally chromium and nickel. Ferralsols typically have resistent, low acticity clays, i.e., kaolins and gibbsite. Ferralsols may contain a wide range of clay contents, however, and in contrast to Ultisols and Alfisols, they have relatively constant clay contents throughout their profiles. Often Ferralsols have clays that form very durable aggregates, a structure that can give them water-holding characteristics that are similar to coarse-textured soils. As a result, even Ferralsols with high clay contents are often drought prone, due to low water holding capacities. Low water holding capacity on a per unit weight or volume basis may be compensated by the fact that many Ferralsols can be deeply rooted by certain plant species. Some Ferralsols are derived from mafic or basic rocks such as basalt with readily weatherable iron-containing minerals that allow such Ferralsols to form in place. More commonly, Ferralsols are formed in previously weathered, transported materials, rather than formed in place from primary materials that are located on stable landscape positions (Lepsch et al, 1977; Macedo and Bryant, 1987; Buol et al., 1989; Buol and Eswaran, 1988). As a consequence, Ferralsols often obtain their pre-weathered mineral characteristics from processes not related to their present locations. The regional distribution of Ferralsols is markedly independent of current regional precipitation patterns, and many Ferral-

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sols support tropical savanna, such as the Cerrado in Brazil (Macedo and Bryant, 1987; Buol et al., 1989; Buol and Eswaran, 1988). A wide range of natural vegetation and ecosystems is supported by Ferralsols, including tropical rainforest, shrub and thorn forests, semi-deciduous forests, savannas, and grasslands. Many of these Ferralsols in semi-arid climates are thus relics, formed mainly in ancient humid environments. Ferralsols may be acidic to neutral in pH, but many have low CEC with a large fraction of variable charged exchange sites (Uehara and Gillman, 1981). Although variable charge may dominate Ferralsols, under field conditions, relatively few appear to have large net positive charge (Cochrane et a1 ., 1985). Nevertheless, relatively low contents for plant-available P and exchangeable Ca, Mg, and K may readily create nutritional problems for plants. On the other hand, physical properties of many Ferralsols make these soils well suited for agronomic use, although many Ferralsols have not been intensively managed over the long-term. Ferralsols are the most prominant of the world’s potentially arable but currently unexploited soils of the world (Brady, 1984). Recent estimates of the Cerrado of Brazil, for example, indicate that about 100 million hectares of Ferralsols could be potentially cultivated and another 100 million hectares converted to pasture (J. Macedo, personal communication). Ferralsols are not uniform in their properties (Tables 2 and 7). In the 1974 FAO/UNESCO system, six secondary mapping units of Ferralsols were recognized (Table 7). These secondary mapping units of the 1974 FAO/UNESCO system represent Ferralsols that have high concentrations of organic matter (Humic Ferralsols), or have different degrees of weathering, mineralogies, and plinthic materials. The secondary classification of the 1974 FAO/UNESCO system differ greatly from the suborder classes of the 1988 Oxisols in the Soil Taxonomy system (Table 2), or from the major subtaxa of Latossolos in recent descriptions of the Brazilian classification system (Carmargo et al., 1986). The major subcategories of Latossolos in the Brazilian system of classification depend on CEC, colour, minerology, drainage, and iron chemistry. Although much soil survey and conceptual work has been devoted to Ferralsols in recent years (Buol and Eswaran, 1988; Duchaufour, 1982; Carmargo et al., 1986), these soil taxa need additional mapping and continued conceptual development (Buol and Eswaran, 1988; Van Wambeke, 1989). According to the FAO/UNESCO data, Ferralsols occupy about 21% of the total tropical land mass, or about 1 billion hectares (Tables 6). Ferralsols occupy markedly different proportions of the FAO/UNESCO maps of tropical South America, Africa, Asia, and Central America; about 40.0, 18.5, 1.8, and 0.2%, respectively. According to the FAO/

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UNESCO map, the majority of Ferralsols cover the Amazon and Zaire River basins. In the FAO/UNESCO map of Africa, enormous mapping units of Ferralsol cover the central African basement, the Zaire River basin, the Zambian highlands, the Mozambique belt, and the eastern half of Madagascar. In the FAO/UNESCO map of South America, enormous mapping units of Ferralsols cover the Guyana Shield, the Cerrado uplands, the Brazilian Shield, the Amazon planalto, and Pleistocene terraces. As we will describe later in this paper, recent systematic soil surveys demonstrate that many of these areas are actually not covered by Ferralsols. To a significant extent, the large proportion of the tropics (21%) that is mapped as Ferralsols by the FAO/UNESCO map is the direct result of archaic ideas about the domination of Latosols in the humid tropics. Sanchez (1976) and Buol et af. (1989) suggest that the FAO/UNESCO map of Ferralsols in South America greatly overestimates their occurance in that continent. As more of the soils of the tropics are surveyed, Ferralsols will be recognized to occupy much less area than was indicated on the FAO/UNESCO maps in which many Ferralsols were mapped mainly by regional climate and vegetation types rather than by soil properties. Based on an analysis of the recent Brazilian soil surveys, later in this paper we hypothesize that as little as 580 million hectares (about 12% of the total tropics) will eventually be mapped as Ferralsols, assuming that the diagnostic criteria of this taxa are not radically modified.

C. Acrisols: the Underestimated Soil Order Acrisols are mainly Ultisols in the Soil Taxonomy system. Acrisols are strongly weathered, acid soils with clay-enriched, illuvial subsoils. Like Ferralsols, Acrisols vary greatly between continents in the proportional area they cover, but are most concentrated in humid regions, especially on slightly younger materials compared to Ferralsols. The 1974 FAO/UNESCO system had five secondary mapping units for Acrisols (Table 7), which reflect differences in weathering and mineralogy, moisture conditions, organic matter content, and presence of plinthic material in subsoils. Major management problems with Acrisols are associated with their acidity. Like Ferralsols, Acrisols have few primary minerals remaining, although secondary clay minerals may be prominant, especially in subsoils. Acrisols are mapped by the FAO/UNESCO as covering 29% of tropical Asia, but only about 12, 8 and 4% of South America, Central America, and tropical Africa, respectively. Major Acrisols mapping units in the FAO/UNESCO map are in southeastern China, the Shan

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Plateau of Burma, the Indochinese and Malaysian peninsulas, the Philippines, and the non-volcanic parts of the Sumatran and Kalimantan Islands of Indonesia. Acrisols are mainly mapped in South America by FAO/UNESCO in the upper Rio Madeira and Araguaia watersheds; in Africa, they are in the west African basement complex and the Inter-Rift Valley in Tanzania. Estimated coverage of Acrisols of tropical regions according to the FAO/UNESCO mapping is about 500 million hectares. The actual extent of Acrisols in the tropics is probably substantially higher than the 500 million hectares estimated by the FAO/UNESCO maps. Because the pedogenic time required for Ferralsol formation has been underestimated in the past, many areas of the humid tropics that have been mapped as Ferralsols are actually Acrisols. Our analyses of the Brazilian Amazon detailed later in the paper, indicate that Acrisols in this region were underestimated by FAO/UNESCO mapping by nearly 100%. The new soil surveys of the Brazilian Amazon include about 133 million hectares of Acrisols, compared with the FAO/UNESCO estimate of about 74 million hectares.

D, Lithosols, Arenosols, and Luvisols: From Extremely Fertile to Infertile, 500 Million Hectares Each Three primary mapping units are mapped by the FAO/UNESCO map as each occupying nearly 500 million hectares, or each about 10% of the total area of the tropics (Table 6). These primary mapping units include Lithosols, Arenosols, and Luvisols. Some of the world's most fertile and infertile soils are included in these soils. Lithosols have no equivalent order in the Soil Taxonomy system, but are grouped as a variety of lithic subgroups of several soil orders. Lithosols are typically shallow 'with frequent rock exposure (< 10-cm depth over solid rock), and are most frequently found in mountainous terrains. Many Lithosols found on hill slopes are highly unstable, a situation that greatly limits development of soil horizons and profiles. No secondary Lithosol mapping units are represented in the FAO/UNESCO system (Table 7), which should not suggest uniformity of soil properties. Chemical and physical properties of Lithosols are strongly influenced by characteristics of parent materials which vary widely on regional and local scales. In contrast to Ferralsols and Acrisols, Lithosols occupy relatively similar proportions of tropical Africa, Central and South America, and Asia, about 10% of each (Table 6). In Africa, major Lithosol mapping units are located in Taoudenni basin in Mauritania, in the Guinean highlands, and throughout the Sahara Desert; in South America, Lithosols are found throughout the Andes mountains. Due to the large area

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of tropical Africa, about half of the tropical world’s Lithosols are mapped in Africa, about 230 million hectares. Arenosols are similar to Psamments in the Soil Taxonomy system, soils that form from coarse, sandy materials. Arenosols may have a variety of subsoils, but all such subsoils have < 15% clay. Arenosols are potentially droughty and have only very weakly developed horizons. These sandy soils are divided into four secondary mapping units to reflect differences in colour, structure, accumulation of illuvial clays, accumulation of sesquioxides, and incipient eluvial E horizons. Most of the tropical Arenosols are found in Africa, where they occupy about 371 million ha or 77% of the total Arenosols mapped by FAO/UNESCO in the tropics. Large mapping units are found throughout the sub-Sahara from Chad to Senegal, and in the Kalahari Desert in southwestern Africa. In contrast, Arenosols are not mapped in Central America, and only 4-5% of tropical Asia and South America is estimated to contain Arenosols. Luvisols are similar to Alfisols in the Soil Taxonomy system, soils with clay-enriched, illuvial subsoils that are generally nutrient-rich and low to moderate in acidity. This major soil taxon is perhaps the most well studied of any major group of soils in the world (Rust, 1983). Luvisols are extremely variable as a primary FAO/UNESCO mapping unit. Eight secondary mapping units are recognized, which account for difference in moisture regimes, calcium carbonate content, minerology, shrink-swell capacity, and sesquioxide concentrations. Many Luvisols are highly fertile agricultural soil in the tropics (Sanchez and Buol, 1975). Although seasonality of rainfall often limits annual crop productivity, due t o high native soil fertility Luvisols often are associated with high human population densities. Luvisols occupy major areas scattered throughout the tropics, between about 8 and 13% of each of the four tropical areas detailed in Table 6. Major areas mapped as Luvisols are in the west African basement complex from Nigeria to Mali; in southeastern Africa from Zimbabwe to Kenya; in eastern and southern regions of the Indian peninsula, much of Sri Lanka; and in scattered areas of southern Mexico, Nigaragua, northeastern Brazil, and Andean valleys of Colombia and Bolivia.

E. Regosols, Yermosols and Cambisols: Weakly Developed Soils Three primary mapping units of the 1974 FAO/UNESCO system have relatively weakly developed soil horizons. The three taxa total about 850 million hectares, with each representing between about 4-7% of the tropical total. Regosols are classified as Orthents and Psamments in the Soil

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Taxonomy system, two suborders of the Entisol order. Regosols are thin soils derived from unconsolidated materials, that lack diagnostic soil horizons. Regosols differ widely in fertility, calcium carbonate content, and temperature regime (Table 7). Regosols are mapped to occupy about 336 million hectares in the tropics, about 10% of tropical South America, about 7% of tropical Africa, but only about 1% of tropical Asia. Major mapping units of Regosols are on the Horn of Africa in Ethiopia and Somalia, and across the Saharan Desert generally north of the Arenosols and Luvisols in the region. Yermosols are similar to Aridisols in Soil Taxonomy. Yermosols are very low organic matter soils of deserts which incipient A horizons (surface soils). Yermosols have one of several kinds of B horizons which are usually composed of non-sodium salt, such as calcium sulphate, or alkaline acumulations such as calcium carbonate. The FAO/UNESCO maps them to cover about 270 million hectares of the tropics. About 80% of the Yermosols in the tropics are found in Africa, mainly across the Saharan Desert, whereas only 0.1% are mapped in South America. Cambisols are classified as Inceptisols in the Soil Taxonomy system. They are highly diverse soils which all have weakly developed B horizons, i .e. cambic horizons, that are distinguished by color, structure, or consistence. Eight secondary mapping units are used in the 1974 FAO/UNESCO system to classify a wide variety of soil conditions that differ in acidity, nutrient content, drainage, sesquioxide content, mineralogy, and organic matter content (Table 7). Cambisols are mapped by the FAO/UNESCO to cover about 243 million hectares of the tropics, about 10% of tropical Asia, but only about 1.5% of tropical South America. Major Cambisol mapping units are in Ethiopia, scattered throughout India, and in western Burma. In sum, according to the FAO/UNESCO data, about 80% of the tropical soil surface is covered by eight primary mapping units: Ferralsols, Acrisols, Lithosols, Arenosols, Luvisols, Regosols, Yermosols, and Cambisols. Each primary unit has considerable variation in chemical and physical properties, and in their potential to support various land uses. This diversity in soil properties is demonstrated by the fact that this 80% of the tropics (4 billion hectares) is mapped by 1974 FAO/UNESCO system with 40 secondary mapping units, even at the 1 5 million scale (Table 7).

F. The Other 1 Billion Hectares: Extreme Variation The other 20% of the tropical land suface, about 1 billion hectares, is mapped by FAO/UNESCO with a total of 14 primary mapping units,

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and a total of 52 secondary mapping units. These soils range widely in their properties, from clayey, base-rich Vertisols with extreme shrinkswell capacities to sandy, nutrient-poor and acidic Podzols. The 14 primary map units range in area from about 200 to 6 million hectares each. Nitosofs are similar to certain high clay Alfisols and Ultisols in the Soil Taxonomy system. They are noted for their well structured clayey subsoils that have good water-holding properties and relatively high phosphorus availability. They may be acidic or relatively neutral. Nitosols are mapped to total about 192 million hectares of the tropics. According to the FAO/UNESCO map, Nitosols are mainly found in the African tropics (Table 6), in Ethiopia, Cameroon, Nigeria, and eastern Zaire. Vertisofs of the FAO/UNESCO system are comparable to Vertisols of the Soil Taxonomy system, soils that are marked by very high activity clay that swell and shrink in response to wetting and drying, respectively. Often wetting and drying cause prolific and deep cracking, as well as major volume changes. Although Vertisols are often potentially fertile, they require specialized management techniques to accomodate their unique physical properties. Nonetheless, Vertisols have become soils with major agricultural producing capability. Vertisols occupy about 180 million hectares of the tropics, and according to the FAO/UNESCO map, Vertisols are most common in Asia and Central America, where they cover 7.8 and 5.9% of the tropical land surface, respectively. Major Vertisol mapping units are in west-central India, in Sudan from Kartoum south along the Nile River, scattered throughout southern Chad, and in Mexico. Gfeysofs are not directly comparable to any order of the Soil Taxonomy system, but are classified as aquic suborders in several different orders. Gleysols are poorly drained, often with seasonally pronounced reducing conditions, due to high water tables. Many have high potential productivity for specific agricultural use. Gleysols are a highly diverse soil taxa, and have seven secondary mapping units in the 1974 FAO/UNESCO system (Table 7). Gleysols may vary in acidity, calcium carbonate contents, organic matter accumulation, rooting depth, and presence of plinthic material. Gleysols average about 3% of the total tropical land surface, and are distributed relatively evenly on each of the four continental areas (Table 6 ) . Major mapping units of Gleysols include floodplains of the Amazon and Orinoco Rivers, along the Niger River on the large saltflats southwest of Timbuktu in Mali, in the Niger River delta in Nigeria, surrounding the confluence of the Zaire and Oubangui Rivers in northeastern Zaire and northern Congo, in the headwaters of the Zambezi River in eastern Angola and western Zambia, at the Mouths of the Ganges in Bangladesh, in the lower

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floodplains of the Irrawaddy River in Burma, and in the lower Mekong River floodplains of Cambodia and Vietnem. Xerosols are most closely similar to Mollic Aridisols, in the Soil Taxonomy system, weakly developed dry soils, with incipient A horizons that contain slightly higher concentrations of organic matter than Yermosols. Xerosols are mapped by the FAO/UNESCO map mainly in tropical semi-arid areas of Africa, mainly in Namibia in southwest Africa and in central Sudan, where they occupy an estimated 106 million hectares. Because of their location in semi-arid regions, Xerosols are not well leached, and calcium sulphate and calcium carbonate often dominate soil horizons in the profile. Fluvisols are not correlated with an order in the Soil Taxonomy system, but are most similar to the suborder Fluvents, alluvial Entisols that range widely in soil properties. Of the 134 million hectares mapped throughout the tropics, most are found in Asia and Africa. Fully 5.6% of tropical Asia is mapped as Fluvisols, and much of this land is extremely fertile, such as the many Eutric and Calcaric Fluvents that are under intensive paddy rice management. Most of the major rivers in tropical Asia, Africa, and America have Fluvents as important mapping units along their floodplain terraces. About 35 million hectares of Eutric Fluvisols (nutrient-rich alluvial soils) are mapped in tropical Africa and South America, and have been identified as areas with a large potential for rice production (Greenland, 1981). Fluvisols are, however, diverse with four secondary map units in the 1974 FAO/UNESCO system. Some Fluvents are extremely nutrient poor and acidic, and others are calcarious or sulphidic (Table 7). Planosols are soils that are widely scattered in several orders of the Soil Taxonomy system, namely Alfisols, Ultisols, Aridisols, and Mollisols. Planosols are soils on nearly level landforms with poor drainage, and distinct and abrupt textural boundaries between A and B horizons, a distinction that has earned them the popular name of duplex soils. Planosols have a wide range in organic matter contents, and they are occasionally acidic, whereas others are affected by salt accumulations in climates with relatively high evaporati0n:precipitation ratios. Soil properties range widely, as indicated by the seven secondary mapping units in the FAO/UNESCO system that are found in the tropics (Table 7). The FAO/UNESCO maps Planosols on about 62 million hectares in the tropics (Table 6). About half the tropical world’s Planosols are mapped in South America as nutrient-rich Eutric Planosols, with the largest mapping units located in central Brazil and eastern Bolivia. Kustunozems are most closely similar to Ustolls in the Soil Taxonomy system, organic-rich soils of grasslands, with low acidity and high base saturation. Some Kastanozems have high fertility. The FAO/UNESCO

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map indicates that the taxon covers about 54 million hectares of tropical lands, about 35 million hectares of which are found in Central America, mainly in central Mexico. Andosols are nearly equivalent to the new Andisol order in the Soil Taxonomy system, soils that are derived from volcanic-ash deposits. Although all Andosols are derived from volcanic materials, mainly from deposited ash layers, they range widely in their physical and chemical properties (Leamy, 1988). Due to high allophane contents, Andosols are typically very low in bulk density and high in organic matter often to great soil depths. Andosols are noted for their pronounced ability to stabilize organic compounds. but also for their potential anion adsorption capacity, especially when acidified. They often have high native fertility and excellent physical properties that make them especially suitable for agriculatural even on very steep slopes. Andosols are especially important in Central America where they occupy 7.1% of the land surface. The FAO/UNESCO map indicates that Andosols cover about 20 million hectares in Central America mainly in southern Mexico, Guatemala, El Salvador, Nicaragua, and Costa Rica. These areas represent nearly half the total 44 million hectares of Andosols mapped throughout the tropics. Histosols are nearly equivalent to Histosols in the Soil Taxonomy system. Histosols are formed from accumulations of organic matter, due to high water tables, restricted drainage, or climatic conditions with extremely high precipitation and low evapotranspiration. Histosols can be high or low in acidity. About 66% of the 31 million hectares of Histosols mapped in the tropics have high acidity. About 25 of the 31 million hectares of tropical Histosols are in Asia, most of which are located in coastal lowlands of the Indonesian islands, especially in the eastern coast of Sumatra and other coastal plains of the Indonesian Islands and along the coastline of the Malaysian peninsula (Whitmore, 1984; Burnham, 1984). Solonchaks are similar to some Aridisols in the Soil Taxonomy system, and are those soils dominated by non-sodium salts. They range widely in their concentration and composition of salts, and in their amount of organic matter and moisture regime. About half of the tropical Solonchaks have excess water problems (high water tables), in addition to high salt concentraions. In central Mauritania and Djibouti in Africa, about 210 thousand hectares of Solonchaks are so high in salts and so impenetrable and toxic to plant roots that their secondary mapping modifier is “takyric”, meaning barren of vegetation. Rendzinas are similar to Rendolls in the Soil Taxonomy system, soils that are relatively shallow, organic-rich, and that directly overlie weathering calcium carbonate. The FAO/UNESCO maps nearly 80% of the

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tropical Rendzinas in Central America. They are mapped on about 18 million hectares, and are mainly found on the Yucatan Peninsula of Mexico. Phaeotems are similar to Udolls and Aquolls, two suborders in the Soil Taxonomy system. Phaeozems are moist to wet soils with deep, extremely organic-rich surface horizons with relatively neutral pH and high base saturation. Some are fertile, and according to the map of the world, they are mainly found in southern Brazil. Nearly 80% of Phaeozems in the tropics are mapped in South America. Solonetzes are sodium-dominated soils that are scattered throughout the Soil Taxonomy system as Aridisols, Alfisols, and Mollisols. Solonetzes may have high or low organic matter. The FAO/UNESCO map includes about 12 million hectares in the tropics, most of which are mapped in tropical Africa, mainly in Chad and Somalia. Podzols are closely similar to the Spodosol order in the Soil Taxonomy system. Podzols are acid, coarse-textured sandy soils with accumulations in the subsoil (spodic horizon) of organic matter, and amorphous iron and aluminium oxides. In order of aerial extent, Podzols are last on this list of primary FAO/UNESCO soil mapping units in the tropics, and are estimated to cover about 6 million hectares according to the FAO/ UNESCO map. The world soil map includes Podzols only on maps of Asia and Africa, with none mapped in tropical South America. Podzols were mapped on Sumatra and Kalimantan in Asia and in Zambia and Angola in Africa. In the 1970s, sandy Podzols in Amazonia received considerable attention (Klinge, 1975; Stark, 1978), attention that appears to be out of proportion to their relatively small areal extent in the Amazon and in the tropics as a whole. The new Brazilian soil survey, however, includes about 13 million hectares of Podsols, approximately 2.8% of the 500 million hectare Brazilian Amazon. In sum, the 1974 FAO/UNESCO classification and map demonstrate that the soil taxa in the tropics have a very wide ranging diversity, and that “tropical soil” is a meaningless concept in describing soils in this region. The 1974 FAO/UNESCO map presents Ferralsols to cover about a billion hectares of tropical land, about 20% of tropics, whereas highly weathered Acrisols are estimated to cover an additional 10% of the tropics. It is critical to appreciate that these estimates of Ferralsols and Acrisols are overestimates and underestimates, respectively, probably on the order of hundreds of millions of hectares each. These major mapping errors largely resulted from reliance on outdated concepts of soil taxonomy. Estimates of the extent of soil taxa throughout the tropics can not be quantified in accurate detail until the soils in the tropics are mapped much more systematically than they are at present. It is clear that

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although the FAO/UNESCO map illustrates the diversity of soils in the tropics as a whole, it contains little evidence about the diversity of soils in the humid tropics. We use recent Brazilian soil survey data and a GIS system to analyse soil taxa and mapping units in the Brazilain Amazon, specifically to evaluate the diversity of soils in the humid tropics.

VI. HOW MUCH AREA IN THE TROPICS IS COVERED BY OXISOLS? RESULTS OF THE FIRST SOIL SURVEYS OF THE AMAZON BASIN Misconceptions about soils in the tropics are probably most firmly held about soils in the humid tropics. Such regions include vast areas of the Amazon River basin, the Zaire River basin, throughout the lowlands of southern and southeastern Asia, much of coastal west Africa, and the coastal plains of Central America. In the last two decades, soil surveys, mapping, and agricultural experiments have proceeded in many of these humid regions. Soil taxonomic concepts have been tested in these regions as well. The most impressive advances have arguably come in Brazil, especially in the Brazilian Amazon basin, an enormous 500 million hectare area that represents about 10% of the tropics as a whole. Of the large areas of South America and Africa where so little is known about the details of soils geography (Figs. 4 and 5), recent advances made by the Brazilian soil surveys are of major significance. We use a geographic information system to analyse results of these new Brazilian soil surveys and maps (EMBRAPA, 1981; Carmargo et a l . , 1986) not only to indicate how the results pertain to soils in Brazil per se, but for what the results suggest about soil diversity throughout the humid tropics.

A. Background The Amazon basin is the world’s largest watershed. It covers nearly 500 million hectares in Brazil alone, provided that the Tocantins drainage of the northern Cerrado is included in eastern Amazonia, an inclusion also made by Bates (1864) in his map of the watershed boundaries of Amazonia. The Amazon is covered by a rich variety of vegetation communities, including wet evergreen forests, seasonal forests from almost evergreen to mainly deciduous, freshwater and brackish swampforests which vary depending on duration and depth of inundation, savannas (campo cerrado) on uplands and on poorly drained lowlands, and tropical Andean montane forests and paramo vegetation above treeline. Its rivers dissect the eastern Andean Mountain chain, cut

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through uplands on old continental platforms, and drain immense flat to undulating alluvium, to flat and frequently flooded lower terraces. Seventeen of its tributaries are over 1500 km in length (Shoumatoff, 1986). A wide variety of rock types which form parent materials for soils are represented within its boundaries. Soil and natural resource information has been very late in coming from the Amazon, not without reason since economic development of the region has been slow. There has until recent years been little economic incentive for soil surveys or evaluations of this enormous area. Following native Indian settlement which was extensive throughout the basin, the quest for gold, religious converts, and political subjects stimulated some agriculatural production and established communites over several centuries. Mining and rubber exploitation brought extensive settlement toward the end of the nineteenth century. By the 1940s, serious efforts began to be made to encourage colonization and agricultural development as population pressures along coastal Brazil and in the Andean highlands became more intense (Cochrane et a l . , 1985). Since the 1940s, settlements and agricultural exploitation in Amazonia have had variable success, a trend that suggests a limited knowledge of the various soil resources and their potentials. In fact, the varying success of settlements emphasizes the importance of understanding the diversity of soil resources in the Amazon and elsewhere in the lowland humid tropics (Moran, 1981). Prior to the 1970s, a major soil study of the Amazon was an exploratory investigation conducted by the Division for Soil Survey and Soil Fertility of the Ministry of Agdculature in Rio de Janeiro with collaboration from US-AID and F A 0 (Beek and Bennema, 1966). Although these soil maps provided a wealth of soil data for soils in otherwise unknown regions, the reliability of these maps was relatively poor compared to what is known today. Even during the FAO/UNESCO mapping project, most of the basin was assigned to a soil reliability class I11 (Fig. 4) indicating that as of the early 1980s, few soils maps of Amazonia were based on any soil observations at all. In retrospect, the FAO/UNESCO map of South America might better have included large mapping units of “little known soil complexes” in the Amazonian basin rather than the enormous and deceptively uniform map units of Ferralsols. The misclassification of more than 100 million hectares as Ferralsol west of the confluence of the Rio Madiera, the Amazon, and Rio Negro will be described subsequently, an error that supports the idea that much of the Amazon River basin might have been better mapped by the FAO/UNESCO with a mapping unit of “little known soil complexes” rather than as Ferralsol. The Brazilian government’s decision in the 1970s to stimulate development of the Amazon led to high quality and systematic natural resouce

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inventories which have supplied a wealth of new data about soils, ecology, vegetation, and geologic resources. For the first time a systematic soil survey has been conducted of the Brazilian Amazon. Compared to maps of the past including the FAO/UNESCO, the new Brazilian surveys and maps present a markedly more diverse perspective of soil taxa in Amazonia.

B. Some Common Properties of Amazonian Soils irior to describing some of the diversity of soils within the Amazon basin, it is worthwhile to briefly describe several properties shared by most or all soil in the Amazon. Two characteristics common to most or all soils of the Amazon (and to most soils of the lowland, humid tropics) are the generally seasonal cycle of soil moisture, and the relatively constant soil temperature regime (Sanchez, 1976; Van Wambeke, 1978). Relatively few places in the tropics have rainfall in excess of evapotranspiration (Et) every month of the year. With Et relatively high, plants may exhaust soil water from great depths during dry seasons, which in some parts of eastern Amazonia may be up to about 5-6 months per annum. A second characteristic shared by all soils in the lowland Amazon is that soil temperature is relatively high, with little seasonal variation. Other soil properties are shared by many but not necessarily all soils of Amazonia. Soils have a range of acidities, but most are acidic and at least somewhat low in nutrient status, in that most soluble nutrients that are products of weathering have been removed from soils by leaching or taken up by plants. During the process of soil formation, losses of nutrients to leaching tend to be balanced by inputs from the decomposition of weatherable minerals (if they exist in the parent materials), a source which can also provide nutrients for plant uptake. However, if soil processes have proceeded to form acidic soils by intense leaching and weathering, or if acid soils are formed from geologic materials containing only inert minerals, soil nutrients can be a finite resource. There are many acid soils in Amazonia, and nutrients under these conditions are not readily resupplied by mineral weathering. As is the case with most if not all acid soils throughout the world, long-term cropping systems that remove significant amounts of nutrients in harvests require inputs of nutrients from organic matter and inorganic fertilizers to prevent decreases in soil nutrient availability. Another characterisitc shared by many upland soils is their relatively limited ability to supply water to plants. Plant-available water storage capacity in upland Oxisols may be less than 10% by volume, and plants without deep roots are susceptible to periodic drought stress. Cochrane et al. (1985) mapped about half of Amazonian soils to have water

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holding capacity <7.5% in the surface metre of soil, and many others to have < 15%. A final characteristic of most Amazonian soils is their relatively great susceptibility to water erosion following disturbance of native vegetation. Rainfall is often intense and erosion can be rapidly accelerated from bare ground following forest cutting for agriculture, pastures, roads, towns, or mines. Management of vegetative cover and establishing control over run-off water is very important to minimize such on-site and off-site problems with water resources.

C. The New Soils Map of the Brazilian Amazon Results of our GIS analyses of soil mapping in the Brazilian Amazon are summarized in Figs. 6, 7, and 8, and Tables 8 and 9. The new systematic soil surveys of Amazonia conducted by EMBRAPA indicate that Oxisols were overestimated considerably in the past, and that Ultisols and 'other soil taxa were underestimated. (In the FAO/UNESCO classification, Ferralsols have been overestimated and Acrisols underestimated; in the Brazilian classification, Latossolos and Podzolicos Vermelho-Amarelo Distroficos have been over- and underestimated, respectively). The new maps reduce the coverage of Oxisols in the Brazilian Amazon from about 67.4% to 39.1% (Tables 8 and 9, Fig. 7). These first surveys and maps based on actual soil data indicate convincingly that soil forming processes have not weathered Amazonian soils as intensively as previously thought. The Podzolicos Vermelho-Amarelo Distroficos (closely similar to the Ultisol soil order) is the soil taxonomic unit with the largest increase in area in the recent Brazilian soil surveys (Tables 8 and 9) Ultisols according to the older FAO/UNESCO maps covered approximately 15% of the Brazilian Amazon (mapped as Acrisols), which increased to nearly 30% in the recent Brazilian soil maps (EMBRAPA, 1981). The largest increases of Ultisol mapping units are concentrated in the upland areas west of the confluence of the Rio Negro, the Amazon, and Rio Madeira (Figs. 7, 8). The enormous region of Ultisols, on the order of 100 million hectares, also extends well into Amazonian Peru and Colombia, and was entirely misclassified as Oxisols by the FAO/ UNESCO map (Beek and Bennema, 1966; FAO/UNESCO, 1971). The sedimentary materials that have produced these Ultisols date from the Tertiary in this part of the western Amazon (RADAM, 1977), and are materials that are apparently not sufficiently weathered to have formed Oxisols. This misclassification probably represents one of the largest soil

SOIL DIVERSITY IN THE TROPICS

37 1

Fig.6 New soil maps of the Brazilian Amazon basin (EMBRAPA, 1981; Carmargo el al., 1986) present a clearer perspective of soil diversity in the humid tropics (Table 9 summarizes individual data). Major changes compared to all previous soil maps include far fewer Oxisols (Latossolos in the Brazillian classification) and many additional soil taxa, especially Ultisols (Podzolicos Vermelho Amarelo Distroficos).

mapping errors made in modern times. The recent mapping and surveys also indicate a modestly greater area of potentially fertile soil taxa. Alfisols (both Luvisols and Eutric Nitosols in the FAO/UNESCO classification) were mapped as only about 1.1% of the Brazilian Amazon by the FAO/UNESCO (1971), whereas the new map includes about a four-fold increase in the areal extent of soils with high base saturation and low acidity (Podzolicos VermelhoAmarelo Eutroficos and Terra Roxa Estruturada in the Brazilian classification). Approximately 18.5 million hectares of such soils are mapped by the recent Brazilian survey (Table 9). Important areal changes have also been estimated in soils mapped with plinthite. Estimated areas of Ultisols with plinthite, Plinthic Acrisol in the FAO/UNESCO map of the world, plus Podzolicos Plinticos in the

Figure 7A.

Figure 7B.

373

SOIL DIVERSITY IN THE TROPICS

Table 8 Soil mapping units of the Brazilian Amazon River basin based on GIS-interpreted 1:s million scale FAO/UNESCO Soil Map of the World (FAO/UNESCO, 1971)

Primary Map Unit

Area 100 km?

Ferralsol

3220

Acrisol

736

Arenosol

282

Gleysol Lithisol

Percent of Basin

Secondary Map Unit

Area 1000 km2

67.4

Acric Orthic Xanthic

34 1 1092 1787

7.14 22.87 37.43

15.4

Orthic Plinthic

366 370

7.67 7.7s

5.91

Ferric

282

5.91

257

5.38

Dystric

257

5.38

189

3.96

-

189

Ferric -

28.7 0.88

0.60 0.02

Percent of Basin

3.96

Luvisol

29.6

0.62

Nitosol

24.3

0.51

Eutric

24.3

0.51

Planosol

12.6

0.26

Eutric Solodic

10.5 2.07

0.22 0.04

Solonchak

22.0

0.46

Gleyic

22.0

0.46

0.04

Dystric

Regosol Total

1.84

1.84

0.04

4774

Brazilian map, are mapped as 7.8 to 4.6% of the Brazilian Amazon, respectively, on the two maps. Areas of Psamments (Arenosols and Solos Arenoquartoso, on the two maps, respectively) have changed slightly from 5.9 to 4.4%. Spodosols (Podzols in both maps) were not mapped in the FAO/UNESCO map of the world, but occupy about 2.8% of the new Brazilian map, or about 13 million hectares.

Fig.7 (Facing page) New soil maps of the Brazilian Amazon highlight how the FAO/UNESCO map greatly overestimated the distribution of Oxisols (Ferralsols or Latossolos). Ferralsols are blackened on the FAO/UNESCO map (Fig. 7A), and Latossolos blackened on the EMBRAPA map (Fig. 7B). Most of the areas mapped as Ferralsol by the FAO/UNESCO (1971) were not based on actual soil data (Fig. 4), whereas the EMBRAPA map is based on systematic soil surveys. Note how the region west of the confluences of the Amazon, Rio Negro, and Rio Madeira was entirely misclassified as Ferralsol by the FAO/UNESCO. Elsewhere in the basin, monolithic mapping units of Ferralsol have given way to more detailed and more diverse mapping units. Tables 8 and 9 quantify these data.

Table 9 Soil mapping units of the Brazilian Amazon River basin based on GIs-interpreted 1 5 million scale Map0 de Solos do Brasil (EMBRAPA, 1981; Carmargo et af., 1986) Primary Mapping Unit

Area loo0 km2

Percent of total

Latossolos

1805.8

39.14

Amarelo Distrofico Vermelho-Amarelo Distrofico Roxo Distrofico e Eutrofico Vermelho-Escuro Distrofico Vermelho-EscuroDistrofico e Eutrofico

Podzolicos

1489.6

32.28

Vermelho-Amarelo Distrofico Vermelho-Amarelo Eutrofico Plintico Distrofico

Lateritas Hidromorticas

322.2

6.98

Solos Gley

292.4

Solos Litolicos

Solos Arenoquartosos Podzol Cambissolos

Secondary Mapping Unit

100 km'

Percent of Total

787.3 923.4 1.0 87.6 6.6

17.06 20.01 0.02 1.10 0.14

1118.4 160.8 210.4

24.24 3.48 4.56

Distrofico (Tb) Distrofica e Eutrofica (Ta) Indiscriminadas

273.6 3.0 45.7

5.93 0.06 0.99

6.34

Distroficos Distroficos e Eutroficos

212.8 79.6

4.61 1.72

207.9

4.51

Distroficos Distrofico e Eutroficos Humicos Distroficos

148.6 56.9 2.4

3.22 1.23 0.05

203.2

4.41

Distroficas Hidromoficas Distroficas

178.1 25.2

0.55

130.2

2.82

19.0 42.1 0.9

0.41 0.91 0.02

130.2

2.82

-

62.0

1.34

Distrofico (Tb) Eutroficos (Tb, Ta) Humico Distrofico (Tb)

Area

3.86

Solos Aluviais

35.5

0.77

Grupamento Indiviso de Solos Terras Roxas Estruturadas

32.9

25.0

0.71 0.54

5.5

0.12

1.8

0.04

4614.1

100.00

Solos Salinos Planossolos Total Primary Units

'

Distroficos e Eutroficos Eutroficos

4.4 31.1

0.10 0.68

-

32.9 8.5 2.4 14.1

0.71 0.18 0.05 0.30

Solochak Indiscriminad Costeiros

2.9 2.6

0.06 0.06

Solodico

1.8

0.04

4614.1

100.00

Eutroficas Similar Distrofica e Eutrofica Similar Eutrofica

Total Secondary Units

Figure 8A.

FAO/UNESCO Acrisol

11

Figure 8B.

SOIL DIVERSITY fN THE TROPICS

377

D. How Much Area in the Tropics is Covered by Oxisols? Based on the results of the recent soil surveys of the Brazilian Amazon, we can approach the question of how much of the tropics is covered by Oxisols. We start by assuming that the estimates of 1.0 to 1.1 billion hectares of Ferralsols in the tropics are too high (Tables 5 and 6). This seems a safe assumption given the FAO/UNESCO method of mapping Ferralsols by typically using climatic and vegetation data rather than soils data, and the outcome of the Brazilian soil survey in relation to the FAO/UNESCO map of the Brazilian Amazon (Tables 8 and 9). A first-order approach is to assume that Oxisols in the remainder of South America (mainly in Bolivia, Colombia, Venezuela, and southern Brazil) and in Africa (mainly in the central African basement and the Zaire River basin) have been overestimated by the FAO/UNESCO project to a similar extent as they have been in the 500-million hectare Brazilian Amazon. This apparently bold assumption may not be unreasonable since boundaries of nearly all Ferralsols in South America and in Africa were drawn using much the same procedures (FAO/UNESCO, 1971, 1977a), relying on very few soil observations and almost entirely on a very general knowledge of climate and vegetation (Figs. 4 and 5 ) . According to the new EMBRAPA (1981) soil map, newly mapped Latossolos actually occupy only about 56% of the total area covered by FAO/UNESCO Ferralsols in the Brazilian Amazon (Tables 8 and 9). Given that the FAO/UNESCO maps of the tropics include about 1.05 billion hectares of Ferralsols, 56% of this total amounts to about 580 million hectares of Ferralsols, or about 12% of the tropics as a whole. If a lower-limit estimates of Oxisols in the tropics is about 580 million hectares, an upper limit might simply be estimated by reducing the FAO/UNESCO total for Ferralsols (1.05 billion hectares, Table 6) by the overestimates apparent solely in the Brazilian Amazon that are based on the new EMBRAPA data. This upper-limit estimate of global Oxisols is about 900 million hectares. Given the manner in which Ferralsols were mapped in the FAO/UNESCO project, however, this is almost certainly an overestimate. Considering the enormous FAO/UNESCO mapping units for Ferralsols in tropical South America and Africa, and the way climate and vegetation information rather than

Fig.8 (Facing page) Soil maps of the Brazilian Amazon highlight how the FAO/UNESCO (1971) map greatly underestimated Ultisols (Acrisol or Podzolicos Vermelho Amarelo Distroficos). Blackened areas on the FAO/UNESCO map are Acrisols (Fig. 8A), and on the EMBRAPA map. Podzolicos Vermelho Amarelo Distroficos are blackened (Fig. 8B). Tables 8 and 9 quantify these data.

378

D. D . RICHTER AND L. I . BABBAR

actual soil data were used in mapping most of these areas (Figs. 4 and 5), the so-called lower-limit estimate of global Oxisol coverage, 580 million hectares, is probably our best estimate of Oxisol coverage at this time. Many additional field surveys combined with continued conceptual progress will be needed to estimate the areal extent of Oxisols with more accuracy.

E. A Realistic Concept of Soil Diversity in the Humid Tropics The new soil surveys of the Brazilian Amazon document important improvements in areal estimates of different soil taxa but they also signal major conceptual changes about soils in the humid tropics. This conceptual change departs from that of a “tropical soil”, the relatively monolithic concept that has been critiqued throughout this paper. A more realistic perspective of the diversity of soils in the tropics accomodates a much wider range of soil conditions and soil taxa found in the intertropical regions as a whole, and in the humid tropics specifically. The FAO/UNESCO map of the world demonstrates that soil taxa in the tropics are numerous and varied (Tables 6 and 7), even when considering the small 1 5 million scale of this map. The FAO/UNESCO map is not, however, able to demonstrate soil diversity within the humid tropics, such as the Amazon or Zaire River basins (Figs 4, 5, 6, 7, 8). The major contribution of the new soil surveys of the Brazilian Amazon is that they demonstrate that even in the humid tropics, soils are not as pedogenically aged or as uniform as previously suggested (Fig. 6). Compared to ideas in the past, soils found in the humid tropics represent a wider range of soils taxonomically, a range that spans substantially more pedogenic time and intensity of weathering. Figure 9 (after Smeck et a l . , 1983) helps describe this change in concepts. The figure illustrates free energy of soil profile horizons as a function of pedogenic time, using a simplified evolution of soils from Entisols. to Inceptisols, Alfisols, Ultisols, and finally to Oxisols (Table 1). In fact, soils of a number of tropical landscapes include Entisols, Inceptisols, Alfisols, Mollisols, Ultisols and Oxisols. The more realistic concept of soils in the tropics must consider that pedogenic time can be easily underestimated. Soil formation diagrams such as that in Figure 9 must also be conceived as being multidimensional, since the composition of geologic materials, landforms, local hydrology, biota and various disturbances all influence soil properties and processes and help insure considerable heterogeneity of soils between and within soil orders.

SOIL DIVERSITY IN THE TROPICS

379

Pedogenic time + Fig. 9 Since the pedogenic time to form Oxisols has often been greatly underestimated in the past (Van Wambeke, 1989), there are many other soil taxa in the humid tropics than previously suspected. This simplified diagram (after Wilding and Drees, 1983) illustrates free energy content of soil profiles of several soil orders as a function of pedogenic time (the duration over which bioclimatic factors operate on parent materials in local landscape positions to form soils). Recent soil surveys (e.g. EMBRAPA, 1981) demonstrate clearly that the pedogenic time required for Oxisol formation must be conceived to range far more widely than previously suspected.

VII. MESO- AND LOCAL-SCALE SOIL VARIATION IN THE TROPICS This paper has emphasized that soil diversity in the tropics is relatively easy to demonstrate using maps with scales of 1 5 million. A very large number of soil taxa are represented in the tropics, soils that contrast greatly in their physical, chemical, and biological properties. Local variations in soil properties are also significant in the tropics and at least a comment is worth making about meso- and micro-scale diversity of soil taxa. Local variation in soil properties is at least as great in the tropics as elsewhere in the world. Soil genetic factors that cause soil meso- and micro-scale variability (geology, geomorphology, biota, hydrology, climate and age) are not qualitatively different between tropical and temperate regions, such that spatial variation might be expected to be lower in the tropics (Lepsch et al., 1977; Moormann and Kang, 1978).

3x0

I). D . RIC'HTFR A N D L . I . B A H B A K

There are as yet? notably few analyses of local spatial variability of soils in the tropics, such as that evaluated by Wilding and Drees (1983). Mausbach et al. (1980), in a study of contrasting pedons, hypothesized that for many soil chemical properties, spatial variation of Vertisols< Mollisols = Alfisols < Entisols = Inceptisols = Ultisols < Spodisols. No Oxisols were included in the study. Anecdotal examples illustrate extreme microvariation of soils in the tropics. Moormann and Kang (1978) and Trangmar et al. (1987) described how growth of upland crops in the tropics was sometimes characterized by substantial unevenness over remarkably short distances, variations attributed to edaphic effects and to edaphic by weather interactions. This local variation in productivity may be very great on recently cleared fields with low fertilizer inputs and with large quantities of forest organic matter that ranges widely in chemical composition (e.g. from burned ashes to decomposing wood and foliage) and in nutrient immobilization and mineralization potentials (Trangmar et al., 1987). In tropical forests or savannas, trees may alter soil properties, especially in surface horizons and in the immediate soil environment adjacent to trunks and roots. Forest-soil, tree-soil, and root-soil studies in temperate forests indicate large inter-specific variation in the effects of trees on soil properties (Gersper and Holowaychuk, 1971); Jenny, 1980; Binkley, 1985; Richter, 1986). Factors such as volume and chemistry of stemflow, nutritional requirements of trees, organic chemistry of litter, and N-fixing abilities affect how different tree species alter soils on micro- and meso-scales. Animals also have profound local effects on variations in soil properties. Large areas of many upland soils in humid and subhumid tropics are mixed by termites, sometimes to many meters in depth (Goudie, 1988). In addition to termites, ants, worms, and other soil fauna can also mix great volumes of soil material in many different soils. Human activities also affect soil heterogeneity on a local scale; old village sites often have accumulations of soil organic matter, phosphorus, nitrogen, and other nutrients associated with human and animal wastes, gardening, and kitchen middens. Old anthropic soils along the Amazon River and its tributaries attest to large native populations who formerly inhabited terraces along the river. Humans are also responsible for fires in many locations in the tropics. Fires can have extremely variable effects on properties of surface soils; effects that are both physical and chemical, and are affected on a wide range of spatial scales. Finally, two impressive examples of microscale soil diversity are described in tropical Asia and Africa. At the International Rice Institute in the Philippines, a 50-m trench through an agricultural field contained soil profiles that were common to 6 of the 10 original orders of Soil Taxonomy (Van Wambeke and Dudal, 1978). In west Africa, agronomic

SOIL DIVERSITY IN T H E TROPICS

38 1

field studies have presented plant breeders with complex analytical problems due to extreme within-field variation in edaphic conditions (Moormann and Kang, 1978). Conventional experimental designs proved inadequate for plant-genetics trials in these fields with soils having extreme local variation in fertility.

VIII. CONCLUSIONS The recent paradigm-change in soil taxonomy has encouraged the collection of soil data, and contributed directly to an increased understanding of soil diversity in the tropics. Soil data have accumulated that make it impossible to ignore the diversity of soils found in the tropics, and in the humid tropics specifically. Even small-scale maps (e.g. the FAO/UNESCO soil map of the world) demonstrate elaborately the diversity of soils of the tropics (Tables 6, 7). Such maps document that tropical Africa, America, and Asia have greatly different distributions of primary and secondary map units. Even at the 1:5 million scale, the FAO/UNESCO soil map of the world includes in the tropics 22 different primary mapping units and 97 different secondary mapping units. Many of these mapping units include soils with a wide range of properties. The idea of “tropical soil” is still invoked frequently, despite good evidence and authoritative statements that attest to the complexity of soil taxonomic distributions in the tropics. Perhaps part of the problem is that few ecological scientists know much about soil taxonomy, or have much appreciation for the scientific difficulties created by the absence of an accepted, global system of soil taxonomy. Isbell (1984) contends that even many pedologists do not understand the fundamental principles and limitations of soil classification. It seems extremely important that at the present time the two most widely used soil classification systems (Soil Taxonomy and FAO/UNESCO) appear to be moving apart in their approach and in how each classifies soil (Van Wambeke, 1989). Both systems actually share many principles and approaches, yet neither will remain static in the years ahead. Whether or not these systems continue to diverge, soil taxonomy should be an important part of broadening the perspectives about soils in the tropics. Why misconceptions about “tropical soil” have persisted for so long is a question that a philosopher of science might well examine with interesting results. To the critical natural scientist, misconceptions have managed to persist due to the enormous challenge of surveying soils on the 5 billion-hectare tropical landscape; an absence of interdisciplinary communication about soils in the tropics; problems associated with too many taxonomies and nomenclatures used to describe soils in the tropics; and the course of the development of soil science itself which

382

D. D. RICHTER A N D L. 1. RABBAR

too often has emphasized the factors of soil formation rather than the quantitative properties of soils themselves. Emphasis of many soil and ecological scientists has been on climatic factors of soil formations, which has diminished the importance of complex interactive effects imposed by parent materials, landforms, biota, and soil age. The FAO/UNESCO map of the world contains little indication about soil diversity in the humid tropics, because soil maps of the lowland humid tropics are among the world’s least reliable. Recent soil surveys of Amazonia (EMBRAPA, 1981) and recent conceptual taxonomic developments (Buol and Eswaran, 1988) have made it relatively easy to criticize the notion of “tropical soil” as it applies to soils in the humid tropics. A fundamental soil mapping principle predicts that as larger and larger scale maps are constructed, increasing heterogeneity of soil map units will be recognized. Since mapping units attempt to represent natural occurrences of soils on the landscape, mapping units always contain some associated soils with properties that are outside the range permitted by the specific unit. As soil maps become more detailed (due to more information and larger map scales), discrepancies between soil cartographic and soil taxonomic units will decrease, and increasing numbers of soil taxa will be represented on soil maps. As soil maps of > 1:50,000 scale continue to be constructed for local human settlements and land use planning, the diversity of soils in the tropics will be increasingly obvious and better understood. In sum, the tropics are covered by a variety of soils which in aggregate are much less homogenous than has often been suggested. In place of the archaic notion of the highly weathered “tropical soil”, soils in the tropics, and specifically soils in the humid tropics, range widely in their properties and in their intensities of weathering. There are far fewer Oxisols and many more Ultisols and other soil taxa than estimated in the past, even in the humid tropics. This is well documented by the first soil surveys of the Brazilian Amazon that have been recently completed. To speak carelessly about “tropical soil” greatly oversimplifies the complexity and diversity of ecosystems in this 5-billion hectare region. As soil diversity at all spatial scales (from regional to microsite) is better documented, understanding and use of tropical ecosystems can only improve.

ACKNOWLEDGEMENTS Many thanks to K. Korfmacher for Geographical Information System management; to Drs S. W. Buol and P. A. Sanchez for critical

SOII. I)IVEI<SIT>' IN T H I . TROPIC'S

3x3

discussions a n d t h e use of their library: t o I . F. Lepsch. M . C r a v o . J . Macedo, a n d J . S. Reynolds f o r field trip discussions of soils and landscapes; t o E. Bornemisza, D. Binkley, N . Christensen. R. G . Healy, P. Heine, M. H u s t o n , K. Lal, D. Livingstone, C. H. Periera, W . M. Post. W. Schlesinger, A . Van W a m b e k e , a n d J . Wright for reviews of t h e manuscript: t o M . Doelle a n d N . Stevens for c o m p u t e r work; t o A . M a c F a d y e n f o r constructive editorial c o m m e n t s ; a n d t o P. Wilson and D. Fourqurean for typing.

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Index

.

A bies 247 Acairlospora. 175 Acer saccharurn, 254 Acer sp., 247 Acrisols Amazon basin, 370. 371 tropical, 359-360, 366 Actinorhizal root nodules, 125, 136 effectiveness, 138- 139 infection specificity, 137-138 types, 136-137 Adianturn, 19 African tropical soils, 352ff, 356-357, 38 1 Acrisols, 359 Arenosols, 361 Ferralsols, 358-359 Fluvisols, 364 Gleysols, 363 Lithosols, 360-361 Luvisols, 361 Nitosols, 363 Podzols, 366 Regosols, 362 Solonchaks, 365 Solonetzes, 366 Xerosols, 364 Yermosols, 362 Agropyroti pungens see Elymus pycnanthus Alfisols, Amazon basin, 371 Allium, 215, 219 Allocasuarina, 138 Ainus, 155 actinorhizal symbiosis, 125 ectomycorrhizal symbiosis, 144 root nodulation, 137. 138 Ainus crispa, 138 Ainus glutinosa. 137, 138, 139 Ainus incana, 138, 139 AInus rubra. 137, 138, 191 Ainus rugosa, 138

Ainus viridis ssp. crispa, 138 Alnus viridis ssp. sinuata, 138 Arnanita muscaria, 144 Amazon basin, 367-368 soil mapping 368-369 Amazon basin soils, 321, 332ff, 367, 378, 382 Ferrakols, 359, 368 Oxisols, 370, 377-378 Podzols, 366 properties, 369-370 soil maps, 370ff water erosion following disturbance, 370 Ambrosia, 243 American tropical soils, 352ff, 356ff, 362, 381 Acrisols, 359 Andosols, 365 Arenosols, 361 Ferralsols, 359 Fluvisols, 364 Gleysols, 363 Kastanosems, 365 Lithosols, 360-361 Luvisols, 361 Phaeozems, 366 Planosols, 364 Podzols, 366 Regosols, 362 Rendzinas, 366 Vertisols, 363 Yermosols, 362 see also Amazon basin soils Anas penelope, 44 Andosols, tropical, 365 Andropogon gerardii, 241 Arbutus rnenziesii, 144, 154 Arctostaphylos uva-ursi, 144 Arenosols. tropical, 361 Argiope bruennichi, 43 Arrnillariella rnellea, 237

39 1

392

INDEX

Asian tropical soils, 352ff. 356ff, 381 Acrisols, 359-360 Arenosols. 361 Cambisols, 362 Fluvisols, 364 Gleysols, 363-364 Histosols, 365 Lithosols, 360 Podzols, 366 Regosols, 362 Vertisols. 363 Asparagus officinalis, 244 Aspleniurn, 19 Associative rhizosphere bacteria, 124, 125, 139ff genotype specific relationships, 140ff plant competition, 150-152 plant growth response, 140ff rhizosphere colonization, 139-140 Aster sp., 245 Aster tripolium, 34, 36, 38 Australian soil classification system, 343 A vena fatira, 23 1 A vena sativa , 23 1 Azospirillum brasilense, 140, 141, 142

Bacillus sp. plant competition effects, 150- 152 plant growth promotion, 141, 142 Bacillus suhtilis, 140 Betula pendulu, 155 Betula sp., 247 mycorrhizal associations, 153 Boletinellus merulioides, 198 Boletus edulis, 144 Bradyrhizobium, 125, 129, 130, 132, 136 Branta berniclu, 44 Brassica oleraceu, 2 I 7 Brussica sp., 217, 218, 243 Brazilian soil classification, 336, 337ff. 343 Breeding experiments, 100ff statistical analysis, 102-103

Culirlris ulbu, 44 Calidris alpina, 2. 44-45 Callosohruchus analis, 95

Callosohruchus rnticulntu.s, 64, 67 clutch size determination, 70 full-sib/half-sib breeding experiments, 103ff genetic and phenotypic componcnts of fitness, 109 life-history, 93ff manipulative life-history experiments, 99- 100 selection experiments, 107- 108 Callosobruchus rhoriesianus. 95 Cullosohruchiis spp., 95 between species life-history comparisons, 95-06 Calluna \wlgaris, 244 Cambisols, tropical, 361-362 Carex coriacea, 223 Carex sp.. 233 Caryu, 247 Casuarinu root nodulation, 137, 138 Cetiococcutn , 207 Cerrado Ferralsols, 358 Charadrius hiaticulu, 44 Chelidoniurn tnajus, 216 Chelidoniurn sp., 220 Chorthippus alhornnrginatiis, 43 Citrus, 147 Claviceps purpurea infection, 2, W f , 50 Clutch size models, 69-70, 74 Community structure, rhizospherc micro-organisms, 153ff Comptonia peregrina, 138. 139 Conifers, ectomycorrhiza symbiosis, 144 rhizosphere micro-organisms in seedling establishment, 154-155 Conocephalus dorsalis, 43 Cycadales, 136 Cyperiis longus, 32 Dicentra sp.. 216 Dolichonubis linc~atus,43 Drosophila rnelanoguster , 80 Life-history selection experiments, 107, 109 Drosophila sp. fitness costs of reproduction, 97. 112 genetic trade-offs, 93-94 Ectomycorrhizas (ECM), 126, 128, 172

INDEX

allelopathic interactions, 244 associated soil organisms, 182, 208ff benefits to plants, 143-144, 229, 233 definitions, 198 dispersal. 179, 182 disturbed soils, 183, 184-185, 192 diversity i n communities, 193 edaphic/ climatic specificity, 187-188, 191 effectiveness of symbiosis, 144- 146 environmental temperature effects, 205 forest ecosystems, 253ff fungi, 143ff, 176 Hartig net, 126, 172, 195, 198, 210, 213 high elevation conifer range. 155 host nutrient uptake, 196. 206 host pathogen resistance, 225, 226 host plant life histories, 248, 249 host plants, 191. 197-198, 212 host specificity, 144 hyphal activity in soils, 206 hyphal anastomosis, 192- 193 mycorrhizosphere effects, 208ff nutrient cycling in natural ecosystems, 236, 238 organic nutrient substrates, 236, 261 plant competition, 1.52, 242 plant hormone production, 196 plant succession, 192, 246, 247 pollution responses, 226-227 population ecology. 191 propagules, 178, 205-206 regulation of formation, 212, 214 root exudate responses, 209 root phenology, 204 root types, 195. 212 soil structure effects, 207, 208 Elaeagnus root nodulation, 137 Elymirs pycnanthus, 35 Endomycorrhizas, 126- 127 End-yrnion non-scriptus, 205 Enterornorpha spp., 44 Entroplzospora, 175 Environmental correlation coefficient, 81 Epipactis helleborine, 237 Ericoid mycorrhizal associations, 172, 199, 251. 253, 256. 261 allelopathic interactions, 244 dependancy, 23 1 fungi, 176

393

host nutrient uptake, 196 organic forms of nutrients, 236 plant hormone production, 196 soil detoxification, 226, 231, 244 Eucalyptus, 192, 226 Euscelis ohsoletiis, 43 Exodermis, rnycorrhizal associations, 210-211 Fagopyrurn esculentum, 224 Fagus grundiflora , 254 Fagus sp., 191, 247 Fagits sylvatica, 247, 254 FAO/UNESCO soil classification, 336. 337ff, 343, 381, 382 system flexibility, 348 FAO/UNESCO world soil map, 346ff, 378 data base quality, 348-349 limits, 347 tropical soils, 356ff Fcrralsols Amazon basin, 368 tropical, 357-359, 366, 377 Festuca rubra, 27, 35 Fitness, 63, 64-65 in Callosobruchus spp., 95 components of, 64-65, 70, 95 conceptual frameworks, 68 definition, 71-72 genetic/phenotypic correlations, 109 heritable variation, 66, 67 life-history theory, 70ff negative pleiotropy, 83 Fluvisols, tropical, 364 Frankia, 125, 136ff, 155 host specific infectivity, 137-138 symbiotic effectiveness, 138-139 Fraxinus sp., 198, 247 Frequency-dependent selection (ESS), 64, 67, 111, 112 Full-sib comparisons, 102 Fusariurn heterosporurn , 42 Fusariurn sp., 225

G matrix equilibrium, 84ff, 91-93, 104ff, l l O f f Gene-density interactions, 90-91 Gene-environment interactions, 88ff, 109 G and E matrices, 91-93, 110

394 Genetic correlation coefficient, 81 Geranium, 229 Gigaspora, 175 Gilbertiodendron dewevrei, 254 Gleysols, tropical, 363-364 Glomus caledonium, 146 Glomus fasciculatum , 186 Glomus mosseae, 146, 217 Glomus sp., 146, 175, 218 Glomus tenue, 146, 186, 190 Glycine max, 131 Gymnascella, 176 Gymnostoma, 138

Haematopus ostralegus, 44 Half-sib comparisons, 102 Halimione portulacoides, 14, 35 Hartig net, 126, 172, 195, 197, 210, 213 Hebeloma cylindrosporum, 176 Hedysarum coronarium, 146 Helianthus occidentalis, 244 Heritability h2 estimates, 81 Hippophae root nodulation, 137 Histosols, tropical, 365 Holcus lanatus, 146, 152, 191, 241 Hymenoscyphus, 176 Hysterangium sp., 236

Juglans nigra , 147 Juncus maritimus, 35

Kandic horizon, 328, 345, 346 Kastanosems, tropical, 364-365 Koeleria pyramidata, 241 K-strategies, 95

Laccaria laccata, 144 Larix eurolepis, 242 Larix spp., 242, 247 Latosols, tropical, 357 Lectins, Rhizobium root hair attachment, 131-132 Leghaemoglobin, 133- 134 Legume root nodulation, 137 bacteria cross-inoculation groups, 129-130

INDEX

Life-history evolution, 63ff breeding experiments to estimate genetic parameters, l00ff conceptual frameworks, 68ff development of theory, 69-70 environmental patchiness, 64, 67, 112 experimental approaches, 93ff experimental manipulations, 97ff, 111 frequency-dependent selection (ESS), 64,67, 111, 112 G matrix equilibrium, 84ff, 91-93, 104ff, llOff gene-density interactions, 90-91 gene-environment interactions, 88ff maintenance of genetic variation, 86-88 multivariate selection model, 64, 83ff, 111 negative genetic correlations, 83ff negative pleiotropy, 83, 111 non-additive genetic effects, 88 norm of reaction and adaptive evolution, 89 optimization approach, 63ff, 68-69, 110-111 options set manipulations, 98-99 options set mapping, 109-1 10 phenotypic correlation studies, 94ff population dynamics, 76-77 quantitative genetics, 64, 65, 67, 68-69, 77ff, 111 selection experiments, 106ff variance generated by new mutation, 86, 88 Life-history theory, development of, 69-70 fitness, 70ff optimal life-history, 72ff Liliurn longiflorum, 30 Lithosols, tropical, 360-361 Lobelia, 237 Loliirm perenne, 141, 152, 241 rhizosphere bacteria, plant competition effects, 147ff. 15Off Lotus corniculatus, 19 Lupinus, 216 Luvisols, tropical, 36 1

Medicago sativa, 134. 147, 217 Medicago sp., 130

INDEX

Mycorrhizal fungi, 175-176 allelopathic interactions, 243-244 associated soil organisms, 182- 183 climatic/ edaphic specificity, 176ff, M f f , 193, 194 dispersal, 179ff disturbed soils, 183-185 as food source in soils, 1838-184 hyphal activity in soils, 206-207 hyphal anastomosis, 192-193 inhibition by root exudates, 216, 217 plant hormone production, 196 plant interactions, 196ff population ecology, 188ff propagules, 176-178 Mycorrhizas, 122, 125 allelopathic interactions, 261 arid ecosystems, 252 benefits/cost to host, 196, 214, 227-228, 232ff chemical root features, 215ff, 260 colonization of root system, 176, 177, 178 definitions, 198ff dependency, 201-202, 220ff, 229, 231-232, 241 ecology, 173ff edaphic/environmental factors, 249ff, 251ff environmental temperature effects, 203, 204-205 experimental systems, 257ff facultative, 201, 202, 220, 243, 246, 256 forest ecosystems, 253ff formation, 196-197, 211ff high latitude sites, 252, 254 host plant compatibility/specificity, 196-198 host plant ecology, 235ff host plant growth hormone production, 214 host plant pathogen resistance, 225-226 host plants, 194ff interplant nutrient transfer, 241 life cycle, 174 mycorrhizosphere effects, 208ff in natural ecosystems, 171ff, 227ff, 235ff, 262ff nitrogen-fixing bacteria associations, 210 nutrient cycling, 235ff, 25.5, 261

395

obligate, 201, 202, 246 phenology, 202ff plant community structure, 155 plant competition, 152-153, 239ff plant distribution, 247ff plant succession, 244ff pollution responses, 226-227 regulation of associations, 211ff, 214-215 root active area index, 224-22.5 root exodermis, 210-211, 213 root exudate influences, 209, 215, 216 root function, 194-196 root nutrient uptake, 206-207 root structural diversity, 194-196, 235 root system properties, 211ff, 221ff, 240, 259-260 saline solis, 251 soil nutrient availability, 220-221, 228-229, 234 soil structure effects, 207-208 taxonomic aspects, 260 tree plantation establishment, 154 types of association, 126-127, 172 water stress, 227 Mycorrhizosphere, 208ff Myrica gale, 138, 139 Myrica pennsylvanica, 138 Myrica root nodulation, 137, 138, 139 Myxotrichum, 176 Negative pleiotropy, 83, 111 Nereis diversicolor , 43, 44 Nitosols, tropical, 363 Nitrogenase, 133 Nitrogen-fixing bacteria associations with mycorrhizas, 210 cyanobacteria, 136 Non-legume root nodules, 136-137 Non-mycorrhizal plants, 199ff, 21 1, 213, 215, 220, 256 allelopathic interactions, 244 chemical features, 215ff distribution, 248-249, 253 mineral nutrient capture, 224 non-functional VAM infections, 218 plant nutrient competition, 240, 243 plant succession, 245 Norm of reaction and adaptive evolution, 89

396

INDEX

Oncopeltus spp., 94 Optimization models clutch size determination, 69-70, 74 development of theory, 69-70 fitness, 64-65, 70ff optimal life-history, 72ff senescence, 70 trade-offs, 70-71, 73, 74 Orchid mycorrhizas, 172, 199, 253, 256 dependency, 231 epiphytes, 249 fungi, 176 influences on other plants, 237 nutrient transfer, 236, 237 ORSTOM soil classification, 336, 337ff, 343 Oxisols Amazon basin, 370, 377-378 tropical, 377-378

Parasponia andersonii, 136 Paxillus involutus, 225, 236 Pennisetum americanum, 146 Phaeozems, tropical, 366 Philaenus spittnarius, 43 Phleum pratense, 146 Phragmites australis, 35 Piceu, 247 Pieris rapae body size-fecunditiy relationship, 66, 67 body size heritability, 65-66, 67 evolutionary theory and, 65ff heritable variation in fitness component, 66 optimization approach to life-history, 66, 67 Pinus ponderosa, 152 Pinus sp., 244, 247 Pinus sylvestris, 242, 247 Pisolithus tinctorius, 145, 184 Pisutn sativurri, 134 Plagiomniiim medium, 22 Planosols, tropical, 364 Plant competition associative rhizosphere bacteria, 147ff, 150-152 mycorrhizal fungi, 152-153 symbiotic bacteria, 147-150 Plant growth-promoting rhizobacteria (PGPR), 125, 126

plant growth promotion, 140ff rhizosphere colonization, 139 Plantago lanceolara , 146 Plantago sp.. 233, 241 Plantation establishment, soil microflora, 154 Podzols, tropical, 366 Polygonatittn , 229 Populus, 253 Prostephaniis truncatits, 109 Pseudotsugu rnenziesii, 152, 154, 191 Pteridiutn aqiiilinum , 244 Puccinellia tnaritimu, 3, 25 competitive interactions, 34ff, 39-40 genetic variation, 26-27 growth/production, 32, 34, 35 population differentiation, 38 Quantitative genetics, life history evolution 77ff assumptions, 79 dimensionless quantities, 80-8 I direct/indirect selection, 82-83 environmental correlation coefficient, 81 genetic correlation coefficient, 81 heritability h 2 , 81 life history trade-offs, 77-78 model, 79-78 Quercus sp., 247

Rana sylvatica , 94 Ranunculus ucris, 146 Regosols, tropical, 361 -362 Rendzinas, tropical, 365-366 Rhizohium, 122, 125. 129ff, 141, 155, 255 activation of nodulation genes by plant signals, 131 attachment to root hairs, 131-132 infectivity-effectiveness relationship, 135-136, 150 lectins, 131-132 plant competition effects, 148ff. 152 selection pressure in legume symbiosis, 135, 136 specificity of infectivity, 129-133 specificity of symbiotic effectiveness, 133-134 Rhizohium legitminosarurn biovar trifolii, 135, 148

INDEX

Rhizobium leguminosarum biovar viceae, 132, 134 Rhizobium meliloti, 134 Rhizoctonia, 176, 237 Rhizopogon sp., 144 Rhizopogon vinicolor, 152 Rhizosphere micro-organisms, 122ff associative bacteria, 124, 125, 139ff, 150-152 biomass, 124 categories, 124, 125 definitions, 123 effects on mycorrhizal associations, 208-209 plant community structure, 153ff plant competition, 147ff plant growth effects, 128-129 plant growth-promoting rhizobacteria (PGPR), 125, 126, 139, 140ff plant interactions, 122 plant specificity, 127ff population size measurement, 122ff symbiotic, 124,125 Root active area index, 224-225 developmental phases, 195 exudates, 208, 209, 215ff function, 194- 195 growth, mycorrhizal associations, 196 structural diversity, 194-196 system architecture, nutrient absorption/competition, 221ff, 239-240 types, 195 Root-colonizing fungi, 199-200 r-strategies, 95 Salicornia sp., 36 Salix nigra, 229 Salix sp., 253 Salsola kali, 220, 244 Sanguinaria canadensis, 216 Scirpus maritimus , 35 Sclerocystis , 175 Scutellospora, 175 Senecio cambrensis, 22 Senescence, evolution of, 70 Sitophilus oryzae, 90, 109 Soil classification, 320-322, 328, 333, 336, 381

397

new paradigm, 342ff soil, forming factors, 340, 343 systems, 336, 337ff zonal soils concept, 340-341 Soil orders, 321 322 Soil Taxonomy, 321, 328, 336, 337ff, 343ff, 348, 381 diagnostic soil horizons, 343, 344 levels, 345 open-endedness, 343, 346 Oxisol criteria, 345-346 Solidago 245 Solonchaks, tropical, 365 Solonetzes, tropical, 366 Spartina alterniflora, 2, 4, 10, 11, 12, 43 Claviceps purpurea infection, 41 cytology, 15, 18-19 growth/production, 31, 34 historical distribution, 13, 14 isoenzyme phenotypes, 19, 22 variation, 22 Spartina anglica, Iff attempts at resynthesis, 18 beneficial/harmful effects, 47-48 C4 photsynthetic pathway, 2, 4, 31-32, 34, 37 Calidris alpina interaction, 44-45 Claviceps purpurea infection, 2, 40ff, 50 competitive interactions, 34, 35, 38, 39ff control, 47ff cytology, 12, 15ff, 22 die-back, 2, 4, 8, 35, 38, 4Sff discovery, 5, 7 dispersal patterns, 27ff ecology, 27ff genetic uniformity, 2, 3, 4, 10, 26, 42 geographical range, 10, 48-49 global warming effects, 4, 34, 49 growth, 30ff historical aspects, 4ff, 10-11 introductions, 8-10 invertebrate interactions, 43-44 isoenzyme phenotypes, 19-20, 25-26 long-term survival, 48ff morphological aspects, 12-13, 23-24 niche, 2, 3, 35ff origin, 1-2, 3, 4, lOff, 22 physiological adaptations, 37 production, 30ff

398

INDEX

Spartina anglica, (contd.) Puccinellia niaritirna competition, 34ff. 39-40 seed proteins, 25, 26 seed set, 6, 24, 29-30 self-incompatibility system, 29-30 species interactions, 2-3 spread, 7ff, 27ff, 49 succession to other communities, 35 UK distribution, 8, 9 variation, 21ff, 50 zonation, 23-24 Spartiria brasiliensis, 10 Spartina glabra, 11, 14 Spartina rnaritirna, 2, 4, 10, 11, 12 cytology, 15, 18-19 historical distribution, 13- 14 isoenzyme phenotypes, 19, 22 variation, 22 Spartina patens, 27 Spartina stricta see Spartina tnaritirria Spartina x neyrautii, 6, 11, 20-21 Spartina x townsendii, 12. 13, 21 attempts at resynthesis, 17-18 Claviceps piirpurea infection, 41 cytology, 15 discovery, 4, 5 distribution, 5-7 isoenzyme phenotypes, 19 spread, 27-28 Siiaeda maritinla, 34 Suillus, 144 Symbiotic rhizosphere micro-organisms, 124, 125 infection specificity, 128

Tolineia rnenziesii, 19 Tragopogon , 2 1, 22 Trifoliin, 131, 132 Trifoliurn incarnaturn, 134 Trifoliutn pratense, 135 Trifoliiirn repens, 30, 132, 133, 134, 135, 141, 147, 152, 153 rhizosphere bacteria, plant competition effects, 147ff, 150ff Trifolium sp., 130, 217 associative rhizosphere bacteria relationships, 141 Rhizobiurn symbiosis, 131-132, 134, 135, 136 Tririga totanus, 44

Triticurn, 147 Tropical soils, 315ff Acrisols, 359-360, 366 Andosols, 365 Arenosols, 361 Cambisols, 361-362 cation exchange capacity, 325, 328-329 characteristics, 325 diversity, 330ff, 343, 350ff, 366-367, 378-379, 381 environmental influences on formation, 317 FAO/UNESCO soil map, 347, 350, 352ff, 356ff Ferralsols, 357-359, 366, 377 Fluvisols, 364 Gleysols, 363-364 Histosols, 365 homogeneity, 325, 329 interdisciplinary miscommunication, 333 irreversible hardening, 325, 329 kandic horizon, 345 Kastanosems, 364-365 laterite soils, 318, 319, 323-324, 325, 329, 342 lateritic soils, 324 Latosols, 325, 342, 357 Lithosols, 360-361 local-scale variation, 379-381 Luvisols, 361 mapping extensive areas, 332-333 misconceptions, 316-317, 319, 322ff, 332, 381, 382 Nitosols, 363 Oxisols, 330-332, 377-378, 367ff Phaeozems, 366 Planosols, 364 Podzols, 366 Regosols, 361-362 Rendzinas, 365-366 soil organic matter, 319, 325, 326-328 Solonchaks, 365 Solonetzes, 366 taxonomic aspects, 317, 320-322, 333, 336, 337ff. 381 Vertisols, 363 weathering, 325, 328 Xerosols, 364 Yermosols, 361-362 Tsuga heterophylla, 191

INDEX

Ulmus, 247 Ultisols. Amazon basin, 370

Vertisols, tropical, 363 Vesicular-arbuscular mycorrhizas (VAM). 126, 128, 146-147, 172 allelopathic interactions, 244 definitions, 198, 199 dispersal, 179 disturbed soils, 183, 184 edaphic/climatic specificity, 185ff environmental temperature effects, 204, 205 forest ecosystems, 253ff fungi, 175-176 high aItitude/latitude sites, 252 host pathogen resestance, 225, 226 host plant specificity, 191. 197 host ranges, 146 hyphal activity in soils, 206 infectivity measurement, 178 interspecific competition, 190 mycorrhizosphere effects, 208ff nutrient cycling in natural ecosystems, 236, 237, 238 pesticides, 226 plant growth response, 147 plant nutrient competition, 152. 241, 242, 243 plant nutrient uptake, 196, 206 plant succession, 245, 246, 247 pollution responses, 226-227

399

population ecology, 188-189 propagules, 176-178. 205 regulation of formation, 212ff root exodermis layer, 210. 21 1 root exudate responses, 209, 216ff root growth, 196 root morphological influences, 195, 212-213 root phenology, 203 salinity tolerance, 227 soil nutrient availability, 221, 229 soil organisms associations. 182 soil properties, 207, 252 spore germination, 209, 216, 217. 218 taxonomic aspects, 186 value to plants in natural ecosystems, 229, 232-233 wetland communities, 251 Vicia f a h a , 134 Vigna unguicirlara, 146 Xerosols. tropical, 364 Yermosols, tropical, 361-362 Zaire River basin soils, 321, 332, 333, 377, 378 Ferralsols, 3.59 Zostera, 44

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Advances in Ecological Research Voumes 1-20 Cumulative List of Titles A century of evolution in Spartina anglica, 21, 1 Aerial heavy metal pollution and terrestrial ecosystems, 11, 218 Analysis of processcs involved in the natural control of insects, 2, 1 Ant-plant-hornoptenin interactions, 16, 53 Biological strategies of nutrient cycling in soil systems. 13, 1 Bray-Curtis ordination: an effective strategy for analysis of multivariate ecological data. 14, I Can a general hypothesis explain population cycles of forest lepidoptera'? 18, 179 communities of parasitoids associated with leafhoppers and planthoppers in Europe, 17, 282 Community structure and interaction webs in shallow marine hard-bottom communities: Tests of an environmental stress model, 19, 189 The decomposition o f emergent niacrophytes in fresh water, 14, 115 Dendroecology: A tool for evaluating variations in past and present forest environments, 19, 11 I Developments in ccophysiological research on soil invertebrates, 16, 175 The direct effects of increase in the global atmospheric C 0 2 concentration on natural and commercial temperate trees and forests, 19, 2 The distrihution and abundance o f lake-dwelling Triclads - towards a hypothesis. 3. 1 The dynamics of aquatic ecosystems 6, 1 The dynamics of field population o f the pine looper, Bupalus piniarius L. (Lep., Geom.), 3 , 207 Earthworm biotechnology, and global biogeochemistry, 15, 379 Ecological aspects of fishery research, 7, 114 Ecological conditions affecting the production of wild herbivorous mammals on grasslands, 6, 137 Ecological implications of dividing plants into groups with distinct photosynthetic production capabilities, 7, 87 Ecological implications o f specificity between plants and rhizosphere micro-organisms, 21, 122 Ecological studies at Lough h e , 4, 198 Ecological studies at Lough Hyne. 17; 115 The ecology of the Cinnabar moth, 12, 1 Ecology o f the coarse woody debris in temperate ecosystems, 15, 133 Ecology, evolution and energetics: a study in metabolic adaptation, 10, 1 Ecology o f fire in grasslands, 5, 209 Ecology of mushroom-feeding Drosophilidae, 20, 225 The ecology of pierid butterflies: dynamics and interactions, 15, 51 The ecology of serpentine soils, 9, 255

402

CUMLATIVE LIST OF TITLES

Ecology, systematics and evolution of Australian frogs, 5, 37 The effects of modern agriculture, nest predation and game management on the population ecology of partridges ( P e r d h perdix and Alectoris nrfu), 11, 2 El Niiio effects on Southern California kelp forest communities, 17, 243 Energetics, terestrial field studies and animal productivity, 3, 73 Energy in animal ecology, 1. 69 Estimating forest growth and efficiency in relation to canopy leaf area. 13. 327 Evolutionary and ecophysiological responses of mountain plants to the growing season environment, 20, 60 The evolutionary consequences of interspecific competition, 12, 127 Forty years of genecology, 2, 159 The general biology and thermal balance of penguins, 4, 131 General ecological principles which are illustrated by population studies of Uropod mites, 19, 304 Genetic and phenotypic aspects of life-history evolution in animals, 21, 63 Geochemical monitoring of atmospheric heavy metal pollution: theory and applications, 18, 65 Heavy metal tolerance in plants, 7. 2 Herbivores and plant tannins, 19, 263 Human ecology as an interdisciplinary concept: a critical inquiry, 8, 2 Industrial melanism and the urban environment, 11, 373 Insect herbivory below ground, 20, 1 Integration. identity and stability in the plant association, 6, 84 Isopods and their terrestrial environment, 17, 188 Landscape ecology as an emerging branch of human ecosystems science, 12, 189 Litter production in forests of the world, 2, 101 Mathematical model building with an application to determine the distribution of DursbanB insecticide added to a simulated ecosystem, 9, 133 The method of successive approximation in descriptive ecology, 1, 35 Mutualistic interactions in freshwater modular systems with molluscan components, 20, 126 Mycorrhizal links between plants: their functioning and ecological significance, 18, 243 Mycorrhizas in natural ecosystems, 21, 171 Nutrient cycles and H + budgets of forest ecosystems, 16, 1 On the evolutionary pathways resulting in C4 photosynthesis and craswlacean acid metabolism (CAM), 19, 58 Pattern and process in competition 4, 1 Phytophages of xylem and phloem: a comparison of animal and plant sap-feeders, 13, 135 The population biology and turbellaria with special reference to the freshwater triclads of the British Isles, 13, 235 Population cycles in small mammals, 8, 268 Population regulation in animals with complex life-histories: formulation and analysis of a damselfly model, 17, 1 Predation and population stability, 9, 1 The pressure chamber as an instrument for ecological research, 9, 165 Principles of predator-prey interaction in theoretical experimental and natural population systems, 16, 249 The production of marine plankton, 3, 117 Production, turnover, and nutrient dynamics of above- and below-ground detritus of world forests. 15. 303

CUMULATIVE LIST OF TITLES

403

Quantitative ecology and the woodland ecosystem concept, 1, 103 Realistic models in population ecology, 8, 200 Renewable energy from plants: bypassing fossilization, 14. 57 Rodent long distance orientation ("homing"), 10, 63 Secondary production in inland waters, 10. 91 The self-thinning rule, 14, 167 A simulation model of animal movement patterns, 6, 185 Soil arthropod sampling, 1, 1 Soil diversity in the tropics, 21, 316 Stomata1 control of transpiration: Scaling up from leaf to region, 15, 1 Structure and function of microphytic soil crusts in wildland ecosystems of arid to semi-arid regions, 20, 180 Studies on the cereal ecosystem. 8. 108 Studies on grassland leafhoppers (Auchenorrhyncha, Homoptera) and their natural enemies, 11, 82 Studies on the insect fauna on Scotch Broom Sarotharnnus scopariirs (L.) Wimmer, 5 . 88 Sunflecks and their importance to forest understorey plants, 18, I A synopsis of the pesticide problem, 4, 75 Theories dealing with the ecology of landbirds on islands, 11, 329 A theory of gradient analysis, 18, 271 Throughfall and stemflow in the forest nutrient cycle, 13, 57 Towards understanding ecosystems, 5 , 1 The use of statistics in phytosociology, 2 , 59 Vegetation, fire and herbivore interactions in heathland, 16, 87 Vegetational distribution, tree growth and crop success in relation to recent climate change, 7, 177 The zonation of plants in freshwater lakes, 12, 37

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