Reviews of Environmental Contamination and Toxicology VOLUME 177
Reviews of Environmental Contamination and Toxicology Continuation of Residue Reviews
Editor
George W. Ware Editorial Board Lilia A. Albert, Xalapa, Veracruz, Mexico F. Bro-Rasmussen, Lyngby, Denmark ⴢ D.G. Crosby, Davis, California, USA Pim de Voogt, Amsterdam, The Netherlands ⴢ H. Frehse, Leverkusen-Bayerwerk, Germany O. Hutzinger, Bayreuth, Germany ⴢ Foster L. Mayer, Gulf Breeze, Florida, USA D.P. Morgan, Cedar Rapids, Iowa, USA ⴢ Douglas L. Park, Washington DC, USA Raymond S.H. Yang, Fort Collins, Colorado, USA Founding Editor Francis A. Gunther
VOLUME 177
Coordinating Board of Editors DR. GEORGE W. WARE, Editor Reviews of Environmental Contamination and Toxicology 5794 E. Camino del Celador Tucson, Arizona 85750, USA (520) 299-3735 (phone and FAX) DR. HERBERT N. NIGG, Editor Bulletin of Environmental Contamination and Toxicology University of Florida 700 Experimental Station Road Lake Alfred, Florida 33850, USA (941) 956-1151; FAX (941) 956-4631 DR. DANIEL R. DOERGE, Editor Archives of Environmental Contamination and Toxicology 6022 Southwind Drive N. Little Rock, Arkansas, 72118, USA (501) 791-3555; FAX (501) 791-2499
Springer-Verlag New York: 175 Fifth Avenue, New York, NY 10010, USA Heidelberg: Postfach 10 52 80, 69042 Heidelberg, Germany Library of Congress Catalog Card Number 62-18595. Printed in the United States of America. ISSN 0179-5953 Printed on acid-free paper. 2003 Springer-Verlag New York, Inc. All rights reserved. This work may not be translated or copied in whole or in part without the written permission of the publisher (Springer-Verlag New York, Inc., 175 Fifth Avenue, New York, NY 10010, USA), except for brief excerpts in connection with reviews or scholarly analysis. Use in connection with any form of information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed is forbidden. The use in this publication of trade names, trademarks, service marks, and similar terms, even if they are not identified as such, is not to be taken as an expression of opinion as to whether or not they are subject to proprietary rights. Printed in the United States of America. ISBN 0-387-00214-6
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Foreword
International concern in scientific, industrial, and governmental communities over traces of xenobiotics in foods and in both abiotic and biotic environments has justified the present triumvirate of specialized publications in this field: comprehensive reviews, rapidly published research papers and progress reports, and archival documentations. These three international publications are integrated and scheduled to provide the coherency essential for nonduplicative and current progress in a field as dynamic and complex as environmental contamination and toxicology. This series is reserved exclusively for the diversified literature on “toxic” chemicals in our food, our feeds, our homes, recreational and working surroundings, our domestic animals, our wildlife and ourselves. Tremendous efforts worldwide have been mobilized to evaluate the nature, presence, magnitude, fate, and toxicology of the chemicals loosed upon the earth. Among the sequelae of this broad new emphasis is an undeniable need for an articulated set of authoritative publications, where one can find the latest important world literature produced by these emerging areas of science together with documentation of pertinent ancillary legislation. Research directors and legislative or administrative advisers do not have the time to scan the escalating number of technical publications that may contain articles important to current responsibility. Rather, these individuals need the background provided by detailed reviews and the assurance that the latest information is made available to them, all with minimal literature searching. Similarly, the scientist assigned or attracted to a new problem is required to glean all literature pertinent to the task, to publish new developments or important new experimental details quickly, to inform others of findings that might alter their own efforts, and eventually to publish all his/her supporting data and conclusions for archival purposes. In the fields of environmental contamination and toxicology, the sum of these concerns and responsibilities is decisively addressed by the uniform, encompassing, and timely publication format of the Springer-Verlag (Heidelberg and New York) triumvirate: Reviews of Environmental Contamination and Toxicology [Vol. 1 through 97 (1962–1986) as Residue Reviews] for detailed review articles concerned with any aspects of chemical contaminants, including pesticides, in the total environment with toxicological considerations and consequences. Bulletin of Environmental Contamination and Toxicology (Vol. 1 in 1966) for rapid publication of short reports of significant advances and discoveries in the fields of air, soil, water, and food contamination and pollution as well as v
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methodology and other disciplines concerned with the introduction, presence, and effects of toxicants in the total environment. Archives of Environmental Contamination and Toxicology (Vol.1 in 1973) for important complete articles emphasizing and describing original experimental or theoretical research work pertaining to the scientific aspects of chemical contaminants in the environment. Manuscripts for Reviews and the Archives are in identical formats and are peer reviewed by scientists in the field for adequacy and value; manuscripts for the Bulletin are also reviewed, but are published by photo-offset from cameraready copy to provide the latest results with minimum delay. The individual editors of these three publications comprise the joint Coordinating Board of Editors with referral within the Board of manuscripts submitted to one publication but deemed by major emphasis or length more suitable for one of the others. Coordinating Board of Editors
Preface
Thanks to our news media, today’s lay person may be familiar with such environmental topics as ozone depletion, global warming, greenhouse effect, nuclear and toxic waste disposal, massive marine oil spills, acid rain resulting from atmospheric SO2 and NOx, contamination of the marine commons, deforestation, radioactive leaks from nuclear power generators, free chlorine and CFC (chlorofluorocarbon) effects on the ozone layer, mad cow disease, pesticide residues in foods, green chemistry or green technology, volatile organic compounds (VOCs), hormone- or endocrine-disrupting chemicals, declining sperm counts, and immune system suppression by pesticides, just to cite a few. Some of the more current, and perhaps less familiar, additions include xenobiotic transport, solute transport, Tiers 1 and 2, USEPA to cabinet status, and zerodischarge. These are only the most prevalent topics of national interest. In more localized settings, residents are faced with leaking underground fuel tanks, movement of nitrates and industrial solvents into groundwater, air pollution and “stay-indoors” alerts in our major cities, radon seepage into homes, poor indoor air quality, chemical spills from overturned railroad tank cars, suspected health effects from living near high-voltage transmission lines, and food contamination by “flesh-eating” bacteria and other fungal or bacterial toxins. It should then come as no surprise that the ‘90s generation is the first of mankind to have become afflicted with chemophobia, the pervasive and acute fear of chemicals. There is abundant evidence, however, that virtually all organic chemicals are degraded or dissipated in our not-so-fragile environment, despite efforts by environmental ethicists and the media to persuade us otherwise. However, for most scientists involved in environmental contaminant reduction, there is indeed room for improvement in all spheres. Environmentalism is the newest global political force, resulting in the emergence of multi-national consortia to control pollution and the evolution of the environmental ethic. Will the new politics of the 21st century be a consortium of technologists and environmentalists or a progressive confrontation? These matters are of genuine concern to governmental agencies and legislative bodies around the world, for many serious chemical incidents have resulted from accidents and improper use. For those who make the decisions about how our planet is managed, there is an ongoing need for continual surveillance and intelligent controls to avoid endangering the environment, the public health, and wildlife. Ensuring safety-
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in-use of the many chemicals involved in our highly industrialized culture is a dynamic challenge, for the old, established materials are continually being displaced by newly developed molecules more acceptable to federal and state regulatory agencies, public health officials, and environmentalists. Adequate safety-in-use evaluations of all chemicals persistent in our air, foodstuffs, and drinking water are not simple matters, and they incorporate the judgments of many individuals highly trained in a variety of complex biological, chemical, food technological, medical, pharmacological, and toxicological disciplines. Reviews of Environmental Contamination and Toxicology continues to serve as an integrating factor both in focusing attention on those matters requiring further study and in collating for variously trained readers current knowledge in specific important areas involved with chemical contaminants in the total environment. Previous volumes of Reviews illustrate these objectives. Because manuscripts are published in the order in which they are received in final form, it may seem that some important aspects of analytical chemistry, bioaccumulation, biochemistry, human and animal medicine, legislation, pharmacology, physiology, regulation, and toxicology have been neglected at times. However, these apparent omissions are recognized, and pertinent manuscripts are in preparation. The field is so very large and the interests in it are so varied that the Editor and the Editorial Board earnestly solicit authors and suggestions of underrepresented topics to make this international book series yet more useful and worthwhile. Reviews of Environmental Contamination and Toxicology attempts to provide concise, critical reviews of timely advances, philosophy, and significant areas of accomplished or needed endeavor in the total field of xenobiotics in any segment of the environment, as well as toxicological implications. These reviews can be either general or specific, but properly they may lie in the domains of analytical chemistry and its methodology, biochemistry, human and animal medicine, legislation, pharmacology, physiology, regulation, and toxicology. Certain affairs in food technology concerned specifically with pesticide and other food-additive problems are also appropriate subjects. Justification for the preparation of any review for this book series is that it deals with some aspect of the many real problems arising from the presence of any foreign chemical in our surroundings. Thus, manuscripts may encompass case studies from any country. Added plant or animal pest-control chemicals or their metabolites that may persist into food and animal feeds are within this scope. Food additives (substances deliberately added to foods for flavor, odor, appearance, and preservation, as well as those inadvertently added during manufacture, packing, distribution, and storage) are also considered suitable review material. Additionally, chemical contamination in any manner of air, water, soil, or plant or animal life is within these objectives and their purview.
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Normally, manuscripts are contributed by invitation, but suggested topics are welcome. Preliminary communication with the Editor is recommended before volunteered review manuscripts are submitted. Tucson, Arizona
G.W.W.
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Table of Contents
Foreword ....................................................................................................... Preface ..........................................................................................................
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Role of Phosphorus in (Im)mobilization and Bioavailability of Heavy Metals in the Soil–Plant System ................................................................. NANTHI S. BOLAN, DOMY C. ADRIANO, AND RAVI NAIDU
1
Environmental Fate of Methyl Bromide as a Soil Fumigant ...................... SCOTT R. YATES, JAY GAN, AND SHARON K. PAPIERNIK
45
Disposal and Degradation of Pesticide Waste ............................................. 123 ALLAN S. FELSOT, KENNETH D. RACKE, AND DENIS J. HAMILTON Index ............................................................................................................. 201
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Springer-Verlag 2003
Rev Environ Contam Toxicol 177:1–44
Role of Phosphorus in (Im)mobilization and Bioavailability of Heavy Metals in the Soil–Plant System Nanthi S. Bolan, Domy C. Adriano, and Ravi Naidu Contents I. Introduction .......................................................................................................... II. Sources of Heavy Metals in Soil Environment .................................................. III. Reactions of Metals in Soils ............................................................................... A. Adsorption ...................................................................................................... B. Complexation .................................................................................................. C. Precipitation .................................................................................................... D. Solid-Phase Speciation ................................................................................... IV. Reactions of Phosphate Compounds in Soils ..................................................... A. Water-Soluble Compounds ............................................................................ B. Water-Insoluble Compounds .......................................................................... V. Mechanisms for (Im)mobilization of Heavy Metals by Phosphate Compounds ................................................................................... A. Phosphate Compounds as a Metal Source ..................................................... B. Physiological Phosphorus-Metal Interactions in Plants ................................. C. Adsorption/Desorption of Metals ................................................................... D. Precipitation of Metals ................................................................................... E. Rhizosphere Modifications ............................................................................. Summary .................................................................................................................... Acknowledgments ...................................................................................................... References ..................................................................................................................
1 4 5 6 7 8 8 10 10 10 12 12 20 21 24 29 34 35 35
I. Introduction The term heavy metal in general includes elements (both metals and metalloids) with an atomic density greater than 6 g cm−3 [with the exception of arsenic (As), boron (B), and selenium (Se)]. This group includes both biologically essential
Communicated by George W. Ware. N.S. Bolan ( ) Soil & Earth Science Group, Massey University, Palmerston North, New Zealand D.C. Adriano University of Georgia, Savannah River Ecology Laboratory, Drawer E, Aiken, SC 29802, USA R. Naidu CSIRO Land and Water, Adelaide, South Australia 5064
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N.S. Bolan, D.C. Adriano, and R. Naidu
[e.g., cobalt (Co), copper (Cu), chromium (Cr), manganese (Mn), and zinc (Zn)] and nonessential [e.g., cadmium (Cd), lead (Pb), and mercury (Hg)] elements. The essential elements (for plant, animal, or human nutrition) are required in low concentrations and hence are known as trace elements or micronutrients. The nonessential metals are phytotoxic and/or zootoxic and are widely known as toxic elements (Adriano 2001). Both groups are toxic to plants, animals, and humans at exorbitant concentrations (Adriano 2001; Alloway 1990). Consequently, metals are extensively researched in the life, agricultural, and environmental sciences. Among the metals, Cd, Cr, Cu, Hg, Pb, and Zn have been intensively studied. Other metals, such as silver (Ag) and tin (Sn), are important especially in the mining environment and around industries that process ores containing these metals (Table 1). Health authorities in many parts of the world are becoming increasingly concerned about the effects of heavy metals on environmental and human health and the potential implications to international trade. For example, Cd accumulating in the offal (mainly kidney and liver) of grazing animals not only makes it unsuitable for human consumption but also imperils the access of offal products to overseas markets (Bramley 1990; Roberts et al. 1994). Similarly, bioaccumulation of Cd in potato, wheat, and rice crops has serious implications for local and international commodity marketing. For these reasons, there is global urgency to ensure that the heavy metal content of foodstuffs produced complies with regulatory standards and compares well with those from other countries. Effective action in the long term will depend on gaining an understanding of the causes of heavy metal accumulation and a proper appreciation of these issues for public health. Because of the ever-increasing production of livestock and poultry products for human consumption, more and more organic wastes from these industries have to be handled so as to abide by environmental regulations, including safe disposal onto land. Large quantities of organic amendments, such as poultry manure compost and biosolids, are used as a source of nutrients and also as a soil conditioner to improve the physical properties and fertility of soils. With increasing demand for safe disposal of wastes generated from agricultural and industrial activities, soil is not only considered as a source of nutrients for plant growth but is also used as a sink for the removal of contaminants from these waste materials (Power and Dick 2000). As land treatment becomes one of the important waste management practices, soil is increasingly being seen as a major source of metals reaching the food chain, mainly through plant uptake and animal transfer. Such waste disposal has led to significant buildup of a wide range of metals, such as Cu, Cr, Pb, Cd, Hg, and Zn, and metalloids, such as As, Cr, and Se. Entry of soilborne metals into the food chain depends on the amount and source of metal input, the properties of the soil (especially soil pH and organic matter), the rate of uptake by plants, and the extent of absorption by grazing animals. Nriagu (1988) stated that “this very profound experiment, in which one billion (109) human guinea pigs are being exposed to undue insults of toxic metals, has yet to receive scientific attention that it clearly deserves.”
Table 1. Sources of heavy metals in soils and their expected ionic species in soil solution.
Metal
Density (g cm−3) 5.73
Cadmium (Cd)
8.64
Chromium (Cr)
7.81
Copper (Cu)
8.96
Lead (Pb) Manganese (Mn)
11.35 7.21
Mercury (Hg)
13.55
Molybdenum (Mo) Nickel (Ni)
10.2 8.90
Zinc (Zn)
7.13
As(III): As(OH)3, AsO3− 3 ; As(V): H2As−4, HAsO2− 4 Cd2+, CdOH+, CdCl−, CdHCO+3, CdSO04 Cr(III): Cr3+, CrO−2, CrOH2+, 2− Cr(OH)−4; Cr(VI): Cr2O2− 7 , CrO4 2+ 2+ Cu (II), Cu (III)
Pb2+, PbOH+, PbCl−, PbHCO+3, PbSO04 Mn2+, MnOH+, MnCl−, MnCO03, MnHCO+3, MnSO04 Hg2+, HgOH+, HgCl02, CH3Hg+, Hg(OH)02 − 0 MoO2− 4 , HMoO4, H2MoO4 2+ + 0 Ni , NiSO4, NiHCO3 NiCO03 Zn2+, ZnSO04, ZnCl+, ZnHCO+3, ZnCO03
Contaminant sources Timber treatment, paints, pesticides, geothermal Electroplating, batteries, fertilizers Timber treatment, leather tanning, pesticides, dyes Fungicides, electrical, paints, pigments, timber treatment, fertilizers, mine tailings Batteries, metal products, preservatives, petrol additives Fertilizer
Toxicitya Toxic to plants, humans, and animals Toxic to plants, humans, and animals Cr (VI) toxic to plants, humans, and animalsb Toxic to plants, humans and animals Toxic to plants, humans, and animals Toxic to plants
Instruments, fumigants, geothermal
Toxic to humans and animals
Fertilizer Alloys, batteries, mine tailings
Toxic to animals Toxic to plants, humans, and animals Toxic to plants
Galvanizing, dyes, paints, timber treatment, fertilizers, mine tailings
Phosphorus (Im)mobilization of Heavy Metals
Arsenic (As)
Ionic species in soil solution
aMost
likely to be observed at elevated concentrations in soils and water. is very mobile and highly toxic; however, Cr(III) is essential in animal and human nutrition and is generally immobile in the environment. Source: Adriano (2001).
bCr(VI)
3
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N.S. Bolan, D.C. Adriano, and R. Naidu
With greater public awareness of the implications of contaminated soils on human and animal health, there has been increasing interest within the scientific community in the examination of the transformation and the fate of metals in soils and the development of technologies to remediate contaminated sites. Unlike organic contaminants, most metals do not undergo microbial or chemical degradation, and the total concentration of these metals in soils persists for a long time after their introduction (Adriano 2001). The mobilization of metals in soils for plant uptake and leaching to groundwater can, however, be minimized by reducing the bioavailability of metals through chemical and biological immobilization. Recently, there has been increasing interest in the immobilization of metals using a range of inorganic compounds, such as lime, phosphate (P) compounds (e.g., apatite rocks), and alkaline waste materials, and organic compounds, such as exceptional quality biosolids (Basta et al. 2001; Knox et al. 2000). Although traditionally investigative efforts have examined the potential value of hydroxyapatite in the immobilization of metals, there is growing interest in assessing the effect of other P compounds on metal transformation in soils. Because most of these P compounds are applied to soils as a nutrient source, some of them are also referred to as P fertilizers in this review. In many countries, increasing amounts of phosphate rocks (PRs) are added directly to soils as a source of phosphorus (P) (Bolan et al. 1990; Rajan et al. 1996). Regular application of P fertilizers has been identified as the main source of heavy metal contamination of soils in these countries (McLaughlin et al. 1996; Roberts et al. 1994). Some of these P fertilizers, which act as a source of heavy metal contamination of agricultural soils, have also been found to act as a sink for the immobilization of these metals (McGowen et al. 2001). Phosphate amendment has often been proposed as a practical remediation option for sites with Pb-contaminated soils (Hettiarachchi et al. 2000). This method is considered to be more economical and less disruptive than the conventional remediation option of soil removal. Application of P compounds to soils, however, can have no effect, induce mobilization, or enhance immobilization of metals, the effect being dependent on the nature of P compound, soil type, and metal species (Knox et al. 2000). Although a number of studies have examined the potential value of P compounds in the immobilization of metals in contaminated soils, there has been no comprehensive review on the mechanisms involved in the P-induced (im)mobilization of these metals. Following a brief overview of the reactions of metals and common P compounds that are used as fertilizer in soils, the review focuses on the mechanisms for the (im)mobilization of metals by P compounds. The practical implications of P compounds on the transformation of metals are discussed in relation to sequestration and phytoavailability of metals in soils.
II. Sources of Heavy Metals in Soil Environment Heavy metals reach the soil environment through both pedogenic (or geogenic) and anthropogenic processes. Most metals occur naturally in soil parent materials, chiefly in forms that are not available for plant uptake. Because of their low
Phosphorus (Im)mobilization of Heavy Metals
5
bioavailability, the metals present in the parent materials are often not available for plant uptake and cause minimum impact to soil organisms. Often the concentrations of metals released into the soil system by the natural pedogenic (or weathering) processes are largely related to the origin and nature of the parent material. Apart from Se (Dhillon and Dhillon 1990) and As (Chakraborty and Saha 1987; Naidu and Skinner 1999), other elements (e.g., Cr, Ni, Pb) derived via geogenic processes have limited impact on soil. Unlike pedogenic inputs, metals added through anthropogenic activities typically have high bioavailability (Naidu et al. 1996a). Anthropogenic activities, primarily associated with industrial processes, manufacturing, and the disposal of domestic and industrial waste materials, are the major source of metal enrichment in soils (Adriano 2001) (see Table 1). Atmospheric pollution from Pb-based petrol is a major issue in many developing countries where there is no constraint on the usage of leaded gasoline. Phosphate fertilizers are considered to be the major source of heavy metal input, especially Cd, in pasture soils in Australia and New Zealand (refer to Section V).
III. Reactions of Metals in Soils
cip
sso
tio
ion
Soluble Complexes
ion
Biomass
ion
zat
li era
n
Mi
Layer Silicate Clays
ch Ex
Soil Solution
pt sor
Ab
e
ang
Ion
n
lut
Soluble “Free” Ions
Groundwater
Di
ita
Leaching to
Pre
Plant
Precipitates
Uptake
Metal ions can be retained in the soil largely by (ad)sorption, precipitation, and complexation reactions (Fig. 1). Sorption is defined as the accumulation of matter at the interface between the aqueous solution phase and a solid adsorbent (Sposito 1984). This process can include ion exchange, formation of surface complexes, precipitation, and diffusion into the solid. In many situations, adsorption is believed to be the precursor for subsequent precipitation, and it is difficult to define the boundary separating adsorption and precipitation pro-
Ad
sor
De
sor
pti
on
pti
on
Humus, Oxides, and Allophane
Fig. 1. Reactions of metals in soils. [Source: Adriano (2002)]
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N.S. Bolan, D.C. Adriano, and R. Naidu
cesses (Corey 1981). The lower the metal solution concentration and the more sites available for adsorption, the more likely that adsorption/desorption processes will determine the soil solution concentration (Tiller 1988). Following input into the soil environment, metals interact with the soil mineral and organic phases. However, the fate of metals in the soil environment is dependent on both soil properties and environmental factors. A. Adsorption Charged solute species (ions) are attracted to the charged soil surface by electrostatic attraction and through the formation of specific bonds (Barrow 1985). Retention of charged solutes by charged surfaces is broadly grouped into specific and nonspecific retention (Bolan et al. 1999a; Sposito 1984). In general terms, nonspecific adsorption is a process in which the charge on the ions balances the charge on the soil particles through electrostatic attraction; in contrast, specific adsorption involves chemical bond formation between the ions and the sorption sites on the soil surface (Sposito 1984). If the nonspecific adsorption process solely controls metal adsorption, then the adsorption capacity of the soil is dictated by its cation-exchange capacity (CEC). However, in many soils the amount of metal sorbed exceeds the CEC of the soils (Bolan et al. 1999a). This observation infers that, in addition to nonspecific adsorption, other processes, such as specific adsorption, precipitation, and complex formation, also contribute to metal retention in soils. Both soil properties and soil solution composition determine the dynamic equilibrium between metals in solution and the soil solid phase. The concentration of metals in soil solution is influenced by the pH (Adriano 2001) and the nature of both organic and inorganic ligands (Bolan et al. 1999b; Harter and Naidu 1995; Naidu et al. 1994; Shuman 1986). The effect of pH values above 6 in lowering free metal ion activities in soils has been attributed to the increase in pH-dependent surface charge on oxides of Fe, Al, and Mn, chelation by organic matter, or precipitation of metal hydroxides (Adriano 2001). The effect of pH on the activity of metals in solution in naturally acidic soils is found to decrease with increasing pH. The gradual decrease in heavy metal activity with increasing pH, especially in variable charge soils, is attributed to increasing CEC (Shuman 1986). In general, both the CEC and the total amount of metal removed from soil solution increase with increasing soil pH (Adriano 2001). Three reasons have been given for the effect of inorganic and organic anions on the adsorption of metals (Naidu et al. 1994). First, anions form complexes with metals, thereby reducing their adsorption onto soil particles. Second, the specific adsorption of ligand anions is likely to increase the negative charge on soil particles, thereby increasing the adsorption of heavy metal cations. And third, specifically sorbed anions, such as phosphate (H2PO−4), strongly compete with heavy metal anions, such as arsenate and selenate, resulting in their desorption. Phosphate-induced metal adsorption/desorption reactions in relation to (im) mobilization of heavy metals are discussed in Section V.
Phosphorus (Im)mobilization of Heavy Metals
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B. Complexation Metals form both inorganic and organic complexes with a range of solutes in soils. A number of studies have examined the effect of inorganic anionic complex formation on the adsorption of Cd2+ by soils (Boekhold et al. 1993; Bolan et al. 1999b; Naidu et al. 1994). Most of these studies have indicated that chloride has often been found to form a complex with Cd2+ as CdCl+, thereby decreasing the adsorption of Cd2+ onto soil particles (Naidu et al. 1994, 1996b). However, when the activity of Cd2+ was corrected for complex formation, the adsorption curves for Cd2+ in nitrate (NO−3) and chloride (Cl−) media coincided. O’Connor et al. (1984) showed that while the presence of Cl− ions decreased 2+ adsorption of Cd2+, sulfate (SO2− 4 ) ions increased Cd adsorption relative to com− parable concentrations of chlorate (ClO4) in three calcareous soils. Cadmiumchloro complexation was identified as the active process reducing Cd2+ retention. 2+ The increased retention in the presence of SO2− 4 was attributed to the low Ca 2+ ion activity available for competition with Cd as a result of the formation of the soluble CaSO04 complex. In contrast to inorganic ligand ions, Haas and Horowitz (1986) found that Cd2+ adsorption by kaolinite, a variable charge mineral, was enhanced by the presence of organic matter, which was attributed to the formation of an adsorbed organic layer on the clay surface. As might be expected, the organic component of soil constituents has a high affinity for metal cations because of the presence of ligands or groups that can form chelates with metals (Harter and Naidu 1995). With increasing pH, the carboxyl, phenolic, alcoholic, and carbonyl functional groups in soil organic matter dissociate, thereby increasing the affinity of ligand ions for metal cations. The general order of affinity for metal cations complexed by organic matter is as follows (Adriano 2001): Cu2+ > Cd2+ > Fe2+ > Pb2+ > Ni2+ > Co2+ > Mn2+ > Zn2+ The extent of metal–organic complex formation however, varies with a number of factors including temperature, steric factors (e.g., geometry), and concentration. All these interactions are controlled by solution pH and ionic strength, nature of the metal species, dominant cation, and inorganic and organic ligands present in the soil solution. Metal interactions involving organic ligands were reported for variable charge oxide surfaces where certain organic ligands were found to enhance the adsorption of Cu2+ and Ag+. An alternate mechanism, which appears to be important in temperate soils, involves metal–ligand complexation in solution and subsequent reduction in cation charge, which probably reduces adsorption (Harter and Naidu 1995). The formation of aqueous complexes of Cd with low molecular weight organic acids (LMWOA) from root exudates is expected to dominate the solution chemistry of Cd in rhizosphere. Based on differential pulse anode stripping voltametric and cation-exchange resin extraction data, the dissolved Cd in soil solutions was found to be almost completely complexed with organic matter (Sauve
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N.S. Bolan, D.C. Adriano, and R. Naidu
et al. 2000). Krishnamurti et al. (1997a) observed significant solubilization of Cd from neutral to slightly acidic soils with 0.1 to 1 mM concentrations of acetic, succinic, oxalic, and citric acids, suggesting that Cd release is related to the stability constant of the Cd-LMWOA complex. C. Precipitation Precipitation appears to be the predominant process of metal immobilization in 2− high-pH soils in the presence of anions, such as SO2− 4 , carbonate (CO3 ), hydrox− − ide (OH ), and H2PO4, especially when the concentration of heavy metal ion is high (Adriano 2001). Metalloids such as Cr and As that form anionic species at field soil pH have been reported to form precipitates with cations, such as Ca2+ (Avudainayagam et al. 2001). Coprecipitation of metals, especially in the presence of iron (Fe) and aluminum (Al) oxyhydroxides, has also been reported and often such interactions lead to significant changes in the surface chemical properties of the substrate. Precipitation as metal phosphates is considered to be one of the primary mechanisms for the P-induced immobilization of metals, especially in substrates containing high concentration of metals; this is discussed in more detail in Section V. Liming is often found to increase the retention of metals. For example, Bolan and Thiyagarajan (2001) have observed an increase in the retention of Cr(III) with an increase in pH due to liming. The pH of the lime-treated soil ranged from 7.18 to 8.04, which coincides with the effective precipitation range for Cr(III) as Cr(OH)3 (Rai et al. 1987). Thus, the increased retention of Cr(III) in the presence of the lime is likely due to the formation of Cr(OH)3. The increase in pH from liming is also likely to increase the negative charge of these variable charge soils, which may have enhanced Cr(III) adsorption. D. Solid-Phase Speciation Fractionation studies are often used to examine the effect of amendments, such as lime, P compounds, and biosolid, on the immobilization of metals. Irrespective of the nature of interaction between the metals and soil colloidal particles, following adsorption metal ions redistribute among organic and mineral soil constituents. Fractionation studies suggest that the majority of the metals are associated with organic matter, Fe and Al oxides, and silicate clay minerals in soils. Factors affecting the distribution of metal among different forms include pH, ionic strength of the soil solution, the solid and solution components and their relative concentration and affinities for the metal, and time (Shuman 1991). A large number of sequential extraction schemes have been used for soils, generally attempting to identify metals held in any of the fractions that include soluble, adsorbed/exchangeable, carbonate-bound, organic-bound, amorphous ferromanganese hydrous oxide-bound, crystalline ferromanganese hydrous oxide-bound, and residual or lattice mineral-bound. The most commonly used schemes are modifications of Tessier et al. (1979).
Phosphorus (Im)mobilization of Heavy Metals
9
Metal fractionations using the sequential extraction techniques have primarily been used to identify the fate of metals applied in sewage sludges and in soils contaminated with smelter and mine drainage wastes (Dudka and Chlopecka 1990; Sposito et al. 1982). These studies suggest that treating the soils with sludges or wastes shifts the solid phases of the metals away from immobile fractions to forms that are potentially more mobile, labile, and bioavailable. For example, Dudka and Chlopecka (1990) found that with sewage sludge application the residual forms of Cd2+, Cu2+, and Zn2+ in soil decreased from 34%–43% to 6%–34%, with a corresponding increase in the readily bioavailable forms. The treatment of metal-contaminated soils with P compounds tends to cause the opposite effect in relation to solid-phase metal fractions (Basta et al. 2001; Seaman et al. 2001). Elaborate sequential extraction schemes have frequently been used to identify the distribution of different species of the metal among the various fractions (Krishnamurti 2000; Ross 1994). However, very few attempts have been made to identify the particular species of the metal that contributes to bioavailability. Using an innovative extraction scheme (Krishnamurti and Naidu 2000; Krishnamurti et al. 1995), which estimates the species associated with metal–organic complexes, the importance of this form in the bioavailability of Cd in native soils has been established. Likewise, the importance of metal–fulvic complexes in the phytoavailability of Cu and Zn has also been documented (Krishnamurti and Naidu 2000). Based on the differential Fourier transform infrared (differential FTIR) spectra of the metal–organic complexes extracted by 0.1 M sodium pyrophosphate extractant, Krishnamurti et al. (1997b) showed that Cd in soils was bonded to the carboxyl and the phenolic groups. A logical approach to minimize plant uptake and subsequent contamination of the food chain is to render the trace metals in the soil immobile. The phytoavailability of the different forms of the solid-phase species generally decreases in this order: soluble > exchangeable/adsorbed > organic-bound > carbonate-bound > ferromanganese hydrous oxide-bound > residual or refractory (i.e., fixed in mineral lattice). Immobilization of metals such as Pb, Zn, and Cd could be achieved by additives, such as zeolites (Chlopecka and Adriano 1997; Gworek 1992; Seaman et al. 2001), apatite (Basta et al. 2001; Ma et al. 1993), Mn oxides (Fu et al. 1991; Hettiarachchi et al. 2000), and clay-hydroxy Al polymers (Mench et al. 1994), which may not produce any detrimental byproduct or alter the physicochemical environment of the soils to affect plant growth. Physiologically based in vitro chemical fractionation schemes are increasingly being used to examine the biovailability of metals (Basta and Gradwohl 2000; Ruby et al. 1996). These schemes include the physiologically based extraction test (PBET), potentially bioavailable sequential extraction (PBASE), and the gastrointestinal (GI) test. These improved tests are capable of predicting the bioavailability of metals for both plant uptake and certain soil organisms.
10
N.S. Bolan, D.C. Adriano, and R. Naidu
IV. Reactions of Phosphate Compounds in Soils The P-induced (im)mobilization of metals depends on the form of P compounds used and their reactions in soils. Phosphate compounds that are used as fertilizer are broadly grouped into water-soluble (fast-release) and water-insoluble (slowrelease) fertilizers (Bolan et al. 1993). The important water-soluble P fertilizers include: single superphosphate (SSP), triple superphosphate (TSP), monoammonium phosphate (MAP), and diammonium phosphate (DAP). The important water-insoluble P fertilizers include phosphate rocks (PRs) and basic slag. Partially acidulated phosphate rocks (PAPR) and superphosphate and reactive rock mixtures (e.g., Longlife super in New Zealand) contain both water-soluble and water-insoluble P components. Monocalcium phosphate (MCP) and ammonium phosphate (AMP) are the principal P components present in superphosphates (SSP and TSP) and ammonium phosphates (MAP and DAP), respectively. It is important to understand the reactions of these P compounds in soils to predict the effect of these on the (im)mobilization of metals. A. Water-Soluble Compounds When superphosphate fertilizers are added to soils, the dissolution of MCP results in the formation of slowly soluble dicalcium phosphate (DCP) with a release of phosphoric acid close to the fertilizer granules (Eq. 1). Phosphoric acid subsequently dissociates into H2PO−4 and hydrogen ions (protons–H+). The protons reduce the pH around the fertilizer granules to a very low level (pH < 2). When ammonium phosphate fertilizers are added to soil, they dissociate into ammonium (NH+4) and H2PO−4 ions. The subsequent oxidation of NH+4 to NO−3 results in the release of protons (Eq. 2). Ca (H2PO4)2 + H2O → CaHPO4. H2O + H3PO4 + 4
− 3
+
NH + 2O2 → NO + 2H + H2O
(1) (2)
The acidic solution around the fertilizer granules dissolves the Fe and Al compounds in the soil, resulting in the adsorption and precipitation of P. The pH around the ammonium phosphate fertilizer granules, however, is unlikely to be as low as that around superphosphate fertilizers, causing less adsorption of H2PO−4 ions. The concentration of plant-available P in soil solution decreases with time of contact of fertilizer granules in soils, the decrease depending on the amount of Fe and Al compounds in the soil. The retention of P by soils decreases the amount of P available for both plant uptake and leaching to groundwater. The acidity generated can also have important implications to the mobilization of metals (Section V). B. Water-Insoluble Compounds The solubility of common P compounds that are added to soils and formed as reaction products is given in Table 2. When insoluble P fertilizers such as PRs are added to soil, they must be dissolved in soils for the P to become plant
Table 2. Equilibrium dissolution reaction and the solubility of common crystalline phosphate compounds in soil.
Chemical formula
Equilibrium dissolution reaction
Calcium dihydrogen phosphate
Ca(H2PO4)2
Ca(H2PO4)2(s) ↔ Ca2+ + 2H2PO4−
CaHPO4
CaHPO4(s) ↔
Tricalcium phosphate
Ca3(PO4)2
Ca3(PO4)2(s) ↔
Hydroxy apatite
Ca10(PO4)6(OH)2
Carbonate apatite
Log Ksp
Solubility (g/100 g)
−1.14
18
−6.6
0.14
−24.0
0.02
Ca10(PO4)6(OH)2(s) ↔ 10Ca2+ + 6PO43− + 2OH−
−55.9
Insoluble
Ca10(PO4)6CO3
Ca10(PO4)6CO3(s) ↔ 10Ca2+ + 6PO43− + CO32−
−108.3
Insoluble
Fluroapatite
Ca10(PO4)6F2
Ca10(PO4)6F2(s) ↔
−110.2
Insoluble
Variscite
AlPO4ⴢ2H2O
AlPO4(s) ↔
−21.0
Insoluble
Strengite
FePO4ⴢ2H2O
FePO4ⴢ2H2O(s) ↔ Fe3+ + PO43− + 2H2O
−26.0
Insoluble
Vivianite
Fe3(PO4)2ⴢ8H2O
Fe3(PO4)2ⴢ8H2O (s) ↔ Fe2+ + 2PO43− + 8H2O
Calcium monohydrogen phosphate
Ca2+
−
+ HPO4
Ca2+
Al3+
+ 2PO4
3−
10Ca2+
+ PO4
3−
+ 6PO4
3−
+
2F−
+ 2H2O
−3.11
Insoluble
Phosphorus (Im)mobilization of Heavy Metals
Phosphate compound
Sources: Whitelaw (2000); Stumm and Morgan (1995); Aylward and Findlay (1994); Snoeyink and Jenkins (1980); Lindsay (1971).
11
12
N.S. Bolan, D.C. Adriano, and R. Naidu
available. Dissolution of PRs is a prerequisite not only for the plant availability of P (Rajan et al. 1996) but also for the immobilization of metals through precipitation as metal phosphates (Laperche and Traina 1998). In soils, PRs dissolve by using the acid produced in the soils (Eq. 3); this is a major reason why PRs are very effective as a nutrient source mainly in acid soils (pH < 6.5) (Bolan et al. 1990) and as a metal-immobilizing agent in acid mine drainage (Evangelou and Zhang 1995). The rate of dissolution also depends on the chemical nature and the particle size of the PR. Dissolution rate increases with decreasing particle size. North Carolina phosphate rock (NCPR from the United States), Sechura PR (from Peru), Gafsa PR (from North Africa), and Chatham rise phosphorite (from New Zealand) are considered to be highly reactive. Once the PR is dissolved, the P released undergoes similar adsorption and precipitation reactions as in the case of soluble P fertilizers. Ca10(PO4)6F2 + 12H+ → 10Ca2+ + 6H2PO−4 + 2F−
(3)
Fertilizer brands such as Microbial Phosphate (New Zealand) and Coastal super (in Australia) contain a small amount of water-soluble P, PR, and elemental sulfur (S0). Microbial oxidation of S0 in these fertilizers releases sulfuric acid, which results in the solubilization of PRs (known as the biosuper effect) (Rajan et al. 1996) and the mobilization of metals (Loser et al. 2001; Schippers and Sand 1998).
V. Mechanisms for (Im)mobilization of Heavy Metals by Phosphate Compounds As indicated earlier, P compounds act both as a source and a sink for heavy metals in soils. As a source they enhance the mobilization of metals and as a sink they induce their immobilization. Phosphate compounds affect the (im)mobilization of metals in the soil–plant system through various processes (Fig. 2), which include P–micronutrient imbalance in plant nutrition, direct metal adsorption by P compounds, phosphate anion-induced metal adsorption and desorption, direct precipitation of metals with solution P as metal phosphates, precipitation through the liming action of PRs, and rhizosphere modification through acidification and mycorrhizal association. Both mobilization and immobilization of metals in soils treated with a range of P compounds have been reported, and the probable mechanisms for the P-induced (im)mobilization of metals reported in the literature are given in Tables 3 and 4. A. Phosphate Compounds as a Metal Source Phosphate compounds contain a range of metals (McLaughlin et al. 1996; Mortvedt 1996; Syers et al. 1986) (Table 5). According to Nriagu (1984), “virtually every known element has been found, at least in trace amounts, in a phosphate mineral.” Phosphate minerals are particularly favored as the host of uranium, thorium, and many other rare elements. Addition of P compounds to soils not
Phosphorus (Im)mobilization of Heavy Metals
Fig.2. Plausible mechanisms by which phosphate compounds enhance the (im)mobilization of metals in soils.
13
14
N.S. Bolan, D.C. Adriano, and R. Naidu
Table 3. Selected references on the mobilization of heavy metals by water-soluble and water-insoluble phosphate compounds.
Phosphate compound
Method of Heavy metal investigationa
Water-soluble phosphate compounds NPK fertilizers
Cd
PB
NPK fertilizers
Cd
PB
SSP
Cd
PB
SSP
Cd
PB
TSP
Cd
PB
TSP
Se
PB
TSP
Cu
PB
Ca(H2PO4)2 Ca(H2PO4)2 KH2PO4
As Mo, B, Se Mo
Proposed primary mechanism
Cd source dissolution Cd source dissolution Cd source dissolution Cd source dissolution Cd source dissolution Competitive adsorption Acidification
TL Desorption TL Desorption AD, TL, PB Desorption
KH2PO4
Mo
AD
Desorption
KH2PO4
Cr
AD
KH2PO4
Se
PB
NaH2PO4
Se
PB
Competitive adsorption Competitive root absorption Desorption
As
CF, TL
NaH2PO4
As
CF
NaH2PO4
As
PB
NaH2PO4
As
PB
(NH4)2HPO4
Cd
CF, PB
Competitive adsorption Competitive adsorption Competitive adsorption Competitive adsorption Cd source dissolution
Reference
Singh and Myhr (1998) He and Singh (1994) Williams and David (1976) Gray et al. (1999) Sparrow et al. (1993) Carter et al. (1972) Timmer and Leydon (1980) Qafoku et al. (1999) Qafoku et al. (2001) Nuenhauserer et al. (2001) Xie et al. (1993); Xie and Mackenzie (1991) Aide and Cummings (1997) Broyer et al. (1972)
Hopper and Parker (1999) Creger and Peryea (1994) Reynolds et al. (1999) Woolson et al. (1973) Livesey and Huang (1981) Loganathan et al. (1996)
Phosphorus (Im)mobilization of Heavy Metals
15
Table 3. (Continued).
Phosphate compound
Method of Heavy metal investigationa
Proposed primary mechanism
NH4H2PO4, Ca(H2PO4)2
As
AD
Competitive adsorption
NH4H2PO4, Ca(H2PO4)2 KH2PO4
As
TL
Mo
AD
Competitive adsorption Competitive adsorption
As
CF
PR
Cd
PB
PR
Cd
PB
PR
Cd
PB
PR
Se
PB
Water-insoluble phosphate compounds Hydroxyapatite
Competitive adsorption Cd source dissolution Cd source dissolution Cd source dissolution Competitive adsorption
Reference Peryea (1991); Peryea and Kammereck (1997) Davenport and Peryea (1991) Barrow (1973)
Biosson et al. (1999) Gray et al. (1999) Loganathan et al. (1996) He and Singh (1994) Carter et al. (1972)
SSP: single superphosphate; TSP: triple superphosphate; PR: phosphate rock. aMethod of investigation is as follows: Chemical: adsorption/desorption (AD); chemical fractionation (CF); solubility diagram (SD); transport/leaching (TL). Biological: phytoavailability bioassay (PB).
only helps to overcome the deficiency of some of the essential trace elements, such as Mo, but also introduces toxic metals, such as Cd (McLaughlin et al. 1996). In this regard, Cd contamination of agricultural soils is of particular concern because this metal reaches the food chain through regular use of Cdcontaining fertilizer materials, such as SSP and TSP. This pathway is one of the main reasons this element has been studied extensively in relation to soil and plant factors affecting its bioavailability Accumulation of Cd in soils through regular fertilizer use has been observed in many countries (see Table 3). For example, in New Zealand and Australia, most of the Cd accumulation in pasture soils has been derived from the use of P fertilizers containing a high Cd concentration. The Cd in most P fertilizers originates mainly from the PRs used for manufacturing P fertilizers. It is important to stress that PRs deposits vary in their Cd content from trace to the parts per million (ppm) range, depending on the source of PR. Thus, manufactured P fertilizers also vary accordingly in their Cd content. The Cd in superphosphates is water soluble, and high-analysis P fertilizers, such as TSP, PAPR, and ammonium phosphates, generally contain lower Cd relative to P.
16
N.S. Bolan, D.C. Adriano, and R. Naidu
Table 4. Selected references on the immobilization of heavy metals by water-soluble and water-insoluble phosphate compounds.
P compound
Method of Heavy metal investigationa
Proposed primary mechanism
Water-soluble phosphate compounds Ca(H2PO4)2 Pb CF, PB, XRD Precipitation as metal phosphates Ca(H2PO4)2 Cd, Zn PB Precipitation as metal phosphates K2HPO4, Pb, Zn, Cd CF, PB Phosphate(NH4)2HPO4 induced adsorption K2HPO4 Pb, Zn, Cd XRD, MB Precipitation as metal phosphates KH2PO4, Cd AD, SD PhosphateCa(H2PO4)2 induced adsorption KH2PO4 Cd AD, SD Phosphateinduced adsorption KH2PO4 Cd, Pb CF Phosphateinduced adsorption K2HPO4 Ni XRD, CF Precipitation as metal phosphates (NH4)2HPO4 Cd, Pb, Zn PBET, TL Precipitation as metal phosphates NO4H2PO4, Pb TL Precipitation as Ca(H2PO4)2 metal phosphates Na2HPO4 Pb XRD, XRF, Precipitation as EMPA metal phosphates Na2HPO4 Pb, Zn SEM, EDX, Precipitation as XRD metal phosphates NaH2PO4 Se PB Desorption Soil P
Pb, Cd
SD
Reference
Hettiarachchi et al. (2000) MacLean (1976)
Pierzynski and Schwab (1993) Pearson et al. (2000)
Bolan et al. (1999b)
Naidu et al. (1994)
Tu et al. (2000)
Pratt et al. (1964)
McGowen et al. (2001)
Davenport and Peryea (1991) Ruby et al. (1994)
Cotter-Howells and Capron (1996)
Hopper and Parker (1999) Precipitation as Santillan-Medrano and metal Jurinak (1975) phosphates
Phosphorus (Im)mobilization of Heavy Metals
17
Table 4. (Continued).
P compound
Method of Heavy metal investigationa
Water-insoluble phosphate compounds Hydroxyapatite Cd Hydroxyapatite
Zn
Hydroxyapatite
Cd
Hydroxyapatite
Pb
Hydroxyapatite
Pb
Hydroxyapatite
Cd, Zn
Hydroxyapatite
Pb
Hydroxyapatite
Zn, Pb, Cu, Cd
Hydroxyapatite
Pb
Hydroxyapatite
Pb
Hydroxyapatite
Cd, Pb, Zn
PR (NCPR)
Pb, Zn, Cd
PR
Pb
PR
Pb, Zn, Cd
CF, PB
Proposed primary mechanism
Cation exchange Adsorption
Reference
Jeanjean et al. (1995)
Chlopecka and Adriano (1996) AD Adsorption, ion Mandjiny et al. (1998) exchange, surface complexation, precipitation CF Precipitation as Berti and Cunningham metal (1997) phosphates SEM, EDX, Precipitation as Laperche and Traina IR, EXAPS metal (1998) phosphates Surface Xu et al. (1994) complexation and coprecipitation AD Precipitation as Xu et al. (1994) metal phosphates CF Precipitation/ Boisson et al. (1999) adsorption/ cation exchange AD Precipitation as Zhang et al. (1997) metal phosphates AD, CF Precipitation as Ma et al. (1993) metal phosphates AD, TCLP Adsorption/ Seaman et al. (2001) precipitation CF, GI, Precipitation as Basta et al. (2001) PBET, PB metal phosphates CF Precipitation as Ma and Rao (1997); Ma metal et al. (1997) phosphates XRD, SEM Adsorption/ Chen et al. (1997b) precipitation
18
N.S. Bolan, D.C. Adriano, and R. Naidu
Table 4. (Continued). Proposed primary mechanism
Method of Heavy metal investigationa
P compound PR
Zn
AD
PR waste clay
Cd
CF, PB
Soil P P rich biosolid
Mn Cd
SD AD, SD
Adsorption/ precipitation Adsorption/ precipitation Precipitation Adsorption/ precipitation
Reference Prasad et al. (2001) Gonzalez et al. (1992) Schwab (1989) Soon (1981)
aMethod of investigation is as follows: Chemical: adsorption/partitioning (AD); chemical fractionation (CF); solubility diagram (SD); transport/leaching (TL); toxicity characteristics leaching procedure (TCLP). Minerological: X-ray diffraction (XRD); X-ray fluorescence (XRF); scanning electron microscopic (SEM); electron microprobe analysis (EMPA); energy-dispersive X-ray analysis (EDX); infrared (IR). Biological: phytoavailability bioassay (PB); gastrointestinal test (GI); physiologically based extraction test (PBET); microfauna/macrofauna bioassay (MB). PR: phosphate rock.
Table 5. Metal concentration in phosphate compounds from various sources. Phosphatea compound GPR NFPR JPR NCPR SPR MPR NIPR APR MIPR CRP IRP SSPb TSPb DAPb
Concentration (mg kg−1) As
Cd
Co
Cu
Zn
Mn
4 7 12 23 5 3 3 7 2
38 3 4 48 11 8 100 12 10 2
3 5 <1 2 3 6 6 4 <1 4 109 77 47 16
15 4 8 9 6 4 8 12 6 5 32 15 49 7.2
393 57 235 400 178 90 1010 560 220 95 187 165 418 112
7 212 5 7 91 151 122 2 2 100 975 890 75 307
32 70 10
Ni
9–51
Pb
Hg
<1– 51 0.4–2.1
962 488 238 195
aPhosphate rocks: GPR, Gafsa phosphate rock, NFPR, North Florida phosphate rock; JPR, Jordan phosphate rock; NCPR, North Carolina phosphate rock, SPR, Sechura phosphate rock; MPR, Mexican phosphate rock; NIPR, Nauru Island phosphate rock; ARP, Arad phosphate rock; MIPR, Makatea Island phosphate rock; CRP, Chatham Rise phosphorite; IPR, Indian phosphate rock. Phosphate fertilizers: SSP, single superphosphate; TSP, triple superphosphate; DAP, diammonium phosphate. bDepends on the phosphate rock source. Sources: Adriano (2001); McLaughlin et al. (1996); Syers et al. (1986).
Phosphorus (Im)mobilization of Heavy Metals
19
Comparison between virgin or native (unfertilized) and agricultural (fertilized) soils has often been used to indicate contamination of soil through agricultural practices. Roberts et al. (1994) conducted a survey of 398 sites throughout New Zealand with 312 farm sites and 86 native sites to a depth of 75 mm (Table 6). They obtained evidence for the enrichment of Cd in pastoral soils, and there was a highly significant correlation between total soil P and total soil Cd across all sites, confirming the role of P fertilizers in soil Cd enrichment. Similar results were also obtained for a range of Australian soils (Table 6), which is not surprising considering the long use history in both New Zealand and Australia of superphosphates manufactured from Nauru Island PR. Nauru superphosphates typically contain 34–69 mg Cd kg−1 (Rothbaum et al. 1986). Similarly, Loganathan et al. (1996) examined the movement and distribution of Cd and P in a pastoral soil amended annually for 10 years with four forms of P fertilizers [SSP, DAP, NCPR, and Jordan PR (JPR)] that varied in their total Cd content. Both total and plant-available Cd concentrations decreased with soil depth. Single superphosphate and NCPR, which had higher Cd content, produced a higher Cd concentration than DAP, JPR, and control treatments. Approximately 93% of the applied Cd was recovered in the top 120 mm of soil, and plant recovery of applied Cd ranged from 1.5% to 4.5%. Although many countries have formulated threshold levels for Cd and other heavy metal accumulation in soils because of the use of municipal sewage
Table 6. Cadmium concentration in unfertilized and fertilized surface soils in Australia and New Zealand. Concentration (mg kg−1) Soil type Australiaa Red-brown earth Red podzolic Krasnozem Alluvial Podzol New Zealandb Alluvial Brown-gray loam Gley Peat Yellow-brown earth Yellow-brown loam Yellow-brown peat Yellow-gray earth aFrom bFrom
Williams and David (1976). Roberts et al. (1994).
Unfertilized
Fertilized
0.055 0.024 0.030 0.14 0.033
0.12 0.085 0.30 0.27 0.34
0.13 0.19 0.24 0.22 0.16 0.23 0.31 0.13
0.16 0.49 0.42 0.69 0.22 0.70 0.75 0.12
20
N.S. Bolan, D.C. Adriano, and R. Naidu
sludge, such limits have not been established from fertilizer use. Based on the threshold level for sewage application (3 mg Cd kg−1 soil), the number of years required that would exceed the threshold level in soil through addition of various sources of P fertilizer is presented in Table 7. This table indicates that although fertilizer addition represents the major source of Cd input to soils, at the normal annual rate of fertilizer input (40 kg P ha−1) to pasture soils, the rate of Cd accumulation appears to be very slow. There have been increasing efforts in reducing the accumulation of Cd in soils through the use of low Cd-containing P fertilizers, achieved by either selective use of PRs with low Cd or treating the PRs during processing to remove Cd. Superphosphate fertilizer manufacturers in many countries are introducing voluntary controls on the Cd content of P fertilizers. For example, the fertilizer industry in New Zealand has achieved its objective of lowering the Cd content in P fertilizers from 340 mg Cd kg−1 P in the 1990s to 280 mg Cd kg−1 P by the year 2000. A number of PRs with low Cd content are available that can be used for the manufacture of P fertilizers, but sources with higher Cd content continue to be used in many countries for practical and economic reasons. Alternatively, because Cd has a low boiling point (BP = 789 °C), it can be removed by calcining the PRs. Most phosphoric acid used in the food industry is manufactured only after the removal of Cd through calcination of the PRs. Calcination of PRs may not become a likely option in the fertilizer industry because it is expensive and calcination decreases the reactivity of PRs, making them less suitable for direct application as a source of P (Ando 1987). B. Physiological Phosphorus–Metal Interactions in Plants Large quantities of P fertilizers are added to agricultural soils to overcome soil P deficiency. Excessive accumulation of P in soils followed by its uptake by plants is likely to interfere with the mobilization and uptake of metals in soils. Table 7. Phosphorus (P) and cadmium (Cd) concentrations in various phosphate fertilizers and the calculated number of years required to exceed the threshold concentration of Cd (3 mg Cd kg−1) in soils as a result of fertilizer application. Concentration Phosphate fertilizer Single superphosphate Triple superphosphate Diammonium phosphate North Carolina phosphate rock Sechura phosphate rock Egyptian phosphate rock Gafsa phosphate rock aAt
P (g kg−1)
Cd (mg kg−1)
Years required to exceed the threshold limita
98 190 200 132 131 130 134
32 70 10 54 12 10 70
166 152 1125 135 614 732 107
an annual fertilizer application rate of 40 kg P ha−1.
Phosphorus (Im)mobilization of Heavy Metals
21
For example, continuous application of soluble P fertilizers to soils low in available Zn has often been shown to induce Zn deficiency in plants. Three different factors may be responsible for Zn–P interactions in soils: (i) dilution of zinc in plants by the increase in growth induced by P fertilizers, (ii) inhibition of Zn uptake by cations (Ca in particular) added with P fertilizers, and (iii) phosphate anion-enhanced Zn adsorption in the soil. Inhibition of Zn translocation from root to shoot and physiological inactivation of Zn within the shoot in the presence of high levels of P have also been attributed to P-induced Zn deficiency (Loneragan et al. 1979). Similarly, lead phosphate precipitates have been identified on the root surface and inside the stem and leaf tissues of maize plants grown in P-rich hydroponic culture and in Agrostis grown in mine waste-contaminated soils treated with high levels of soluble P. Applications of lead acetate solutions to plant cells have been shown to result in the formation of tiny crystals of lead pyromorphite within these cells (Cotter-Howells and Capron 1996; Malone et al. 1974). Uptake of certain oxyanions, such as selenite and arsenate, can be inhibited by phosphate because of competitive absorption. Khattak et al. (1991) found that a fourfold increase in P concentration (from 32 to 129 mM) in solution culture caused a 25% reduction in shoot Se in alfalfa. Similarly, Hopper and Parker (1999) observed a decrease in the uptake of Se by ryegrass and clover in the presence of high levels of P. However, they found that sulfate–selenate antagonism was stronger than phosphate–selenate antagonism. Plant response to As and P interactions in the root zone appears to be system dependent. In nutrient solution culture, increasing amounts of soluble P at fixed As concentration reduced As phytotoxicity (Rumberg et al. 1960); this was attributed to the inhibition of As uptake by roots in the presence of excessive amounts of P. In soil systems, however, P fertilizer addition has been shown to enhance the uptake of As because of an increase in the desorption and solubility of As resulting from competitive adsorption of H2PO−4 (Creger and Peryea 1994; Peryea 1991). C. Adsorption/Desorption of Metals Depending on the source, application of P compounds can cause direct adsorption of metals onto these P compounds, enhance anion-induced metal adsorption through increased surface charge, decrease metal adsorption through metal complex formation, and induce desorption of oxyanions, such as chromate and arsenate, through competition (see Table 3). The evidence for these processes is presented next. Direct Adsorption by Apatite Adsorption of metals by apatite mineral has been observed in a number of studies (see Table 4), and the order of heavy metal adsorption by apatite largely depends on pH (Chen et al. 1997a). Adsorption of metals onto hydroxyapatite surfaces has been observed for Zn (Brudevold et al. 1963), Cd (Middelburg and Comans 1991), Sr (Lazic and Vukovic 1991), and
22
N.S. Bolan, D.C. Adriano, and R. Naidu
Ni and Cu (Misra et al. 1975). Similarly, Seaman et al. (2001) observed that the addition of apatite decreased the concentration of a suite of metals including Ni in contaminated pond sediments. Metal adsorption onto apatite is facilitated through the exchange of Ca2+ from the apatite particle by the metal in the soil solution (Suzuki et al. 1981). However, Xu et al. (1994) concluded that in the case of Zn and Cd, surface complexation and coprecipitation (Eq. 4) are the most important mechanisms, with ion exchange and solid-state diffusion also possibly contributing to the overall adsorption process by hydroxyapatite. Ca10(PO4)6(OH)2 + xCd2+ → (Cdx,Ca10−x)(PO4)6(OH)2 + xCa2+
(4)
Phosphate-Induced Metal Adsorption Anion-induced metal adsorption has been reported for a number of cations (Bolland et al. 1977; Ryden and Syers 1976; Shuman 1986; Wann and Uehara 1978a). Ayers and Hagihara (1953) and Wann and Uehara (1978b) showed that leaching losses of K+ in variable charge soils could be reduced by prior application of P fertilizer to the soil. Ryden and Syers (1976) concluded that the retention of Ca2+ in response to H2PO−4 adsorption by soils results from the increase in negative charge induced by H2PO−4 adsorption. Similarly, in examining the effect of inorganic ligands on Cd2+ adsorption by an Oxisol and a Xeralf, Naidu et al. (1994) found that adsorption increased mark− edly in the presence of SO2− 4 and H2PO4 ions. A subsequent study using two soils varying in their variable charge components indicated an insignificant effect of increasing P addition on Cd2+ retention by the soil dominated by permanent charge silicate clay minerals. However, increasing addition of P caused significant increases in the adsorption of Cd2+ by a soil dominated by variable charge components (Bolan et al. 1999b). It has also been shown that Zn2+, Cd2+, or Cu2+ adsorption by variable charge components, such as Al and Fe oxides, and by soils can be enhanced by low or moderate enrichment of oxides with P (Agbenin 1998; Barrow 1987; Bolland et al. 1977; Krishnamurti et al. 1999; Kuo 1986), causing increases in surface negative charge. Several mechanisms can be advanced for H2PO−4-induced metal adsorption by soils, including (i) increase in negative charge, (ii) cosorption of H2PO−4 and metal as an ion pair, and (iii) surface complex formation of metal on the P compound. Although enhanced sorption of a number of metals via increased surface charge in the presence of H2PO−4 ions has been demonstrated (Bolan et al. 1999a), much work is still needed to provide conclusive evidence for mechanisms (ii) and (iii). Studies have shown that specific adsorption of anions increases the net negative charge of variable charge surfaces (Eq. 5) (Table 8). The amount of surface charge acquired through specific adsorption depends on the nature of anion adsorbed, pH, and electrolyte concentration of the solute (Bolan et al. 1999a). The most frequently offered causes for P-induced increase in negative charge (Hingston 1981) include (i) a shift in the zero point of charge to lower pH values, (ii) neutralization of positive charge, and (iii) electrolyte inhibition. − − FeOOH(s) + HPO2− 4 ↔ FeOHPO4(s) + OH
(5)
Phosphorus (Im)mobilization of Heavy Metals
23
Table 8. Observed increases in surface charges of hydrous oxides and soils resulting from specific adsorption of anions.
Soil constituents
Solute
Iron hydrous oxide Phosphate Allophane Phosphate Soil Phosphate Sulfate Soil Phosphate
Soil Soil Aluminum oxide Soil
Phosphate Phosphate Sulfate Sulfate
pH
Charge added (mol mol−1 anion)
6.5 5.1 6.5 6.5 5.0 6.5 7.5 7.0 5.8 5.0 5.6
1.25 0.5 0.65 0.26 0.38 0.47 0.77 0.35–0.7 0.52 1.06 0.25
Reference Bolan et al. (1985) Rajan et al. (1974) Bolan et al. (1986) Sawhney (1974)
Schalscha et al. (1974) Naidu et al. (1990) Rajan (1978) Curtin and Syers (1990)
Bolan et al. (1999b) obtained a good relationship between the increase in negative charge resulting from H2PO−4 enrichment and the increase in Cd2+ adsorption. However, only a small fraction of the increased negative charge was occupied by Cd2+. In their study, when P was added as KH2PO4, 29% of the increase in negative charge due to H2PO−4 adsorption was evidently balanced by Cd2+ adsorption and the remaining 71% by K+ adsorption. In contrast, when P was added as Ca(H2PO4)2, only 8% of the increase in negative charge was balanced by Cd2+ adsorption, the remaining 92% by Ca2+ adsorption. These results indicate that there was a greater competition of Ca2+ for adsorption sites than of K+, as also confirmed by Boekhold et al. (1993). Similarly, Naidu et al. (1994) demonstrated that the effect of ionic strength on Cd2+ adsorption was induced through its effect on surface charge. Addition of specifically adsorbed anions such as phosphate and silicate to soils has been attempted with a view to increasing the negative charge (or cation-exchange capacity, CEC) of variable charge soils (Naidu et al. 1990; Schalscha et al. 1974; Wann and Uehara 1978a). Wann and Uehara (1978a) suggested that addition of P compounds to soils may not only serve as a nutrient but also as an amendment to increase the CEC of the soil. Although P fertilizer application has been considered as a management tool to increase the CEC of variable charge soils, large quantities of fertilizer are required to cause significant increases in CEC. At a maintenance application rate of 40 kg P ha−1, the estimated increase in CEC from such application ranges from 0.07 to 0.18 cmol kg−1 soil (assuming soil bulk density = 1.0 Mg m−3, depth of incorporation of fertilizer = 50 mm, and the increase in surface charge due to H2PO−4 adsorption 0.31–0.70 mol(−) mol P−1).
24
N.S. Bolan, D.C. Adriano, and R. Naidu
Bolland et al. (1977) and Helyar et al. (1976) proposed surface complex formation for the increased adsorption of Ca2+ and Zn2+ onto H2PO−4-enriched gibbsite and goethite. In this case, specifically sorbed anions, such as H2PO−4 form complexes with the soil surface so that cations are adsorbed onto the adsorbed anion. Indeed, surface complexation has been suggested as a mechanism for the immobilization of metals such as Cd and Zn by hydroxyapatite (Xu et al. 1994). Phosphate-Induced Metal Desorption Phosphate anions are very effective in desorbing certain oxyanions, such as selenite, arsenate, and chromate, and phosphate solution is often used as an extractant to measure the amount of adsorbed oxyanions (Aide and Cummings 1997; James et al. 1995). Addition of P fertilizers to lead arsenate-contaminated soils has resulted in an increase in the solubilization and subsequent mobility of As in soils (Davenport and Peryea 1991; Peryea 1991). Seaman et al. (2001) have observed that increasing levels of hydroxyapatite addition to metal-contaminated sediments resulted in increases in the concentrations of chromate and arsenate in solution; this was attributed to increased adsorption competition by H2PO−4 ions with the oxyanions (Barrow 1973; Bartlett and Kimble 1976; Peryea 1991). However, the effect of H2PO−4 on metal desorption depends on the nature of the soil, its sorption capacity, and the extent of saturation with a particular metal ion (Smith et al. 2002). Phosphate has also been shown to strongly compete with molybdate anion (MoO2−4) for adsorption sites, thereby resulting in increased desorption of the latter (Barrow 1973; Xie et al. 1993). Recently, Nuenhauserer et al. (2001) demonstrated the value of P fertilizers in the phytoremediation of Mo-contaminated soils, where application of soluble P fertilizer greatly enhanced the solubilization of Mo in a soil contaminated with excess Mo, facilitating its removal through plant uptake. D. Precipitation of Metals Precipitation as metal phosphates has been proved to be one of the main mechanisms for the immobilization of metals such as Pb and Zn (see Table 4). These new metal phosphate compounds have extremely low solubility over a wide pH range, which makes P application an attractive technology for managing metalcontaminated soils. The formation of the new solid phase (i.e., precipitates) occurs when the ionic product in the solution exceeds the solubility product of that phase. Recent studies using X-ray absorption fine structure spectroscopy (XFAS) indicate that formation of surface precipitates may occur even when solution concentration is undersaturated with respect to homogeneous precipitation of pure metal precipitate phase (Ford and Sparks 2000). In normal soils, precipitation of metals is unlikely, but in highly metal-contaminated soils, this process can play a major role in the immobilization of metals. The solubility of various metal phosphates is given in Table 9. As Metal Phosphates The phosphate radical can combine with more than 30 elements to form phosphate minerals (Nriagu 1984). In addition, a large number
Phosphorus (Im)mobilization of Heavy Metals
25
Table 9. Solubility of common crystalline metal phosphate compounds. Element
Metal phosphate
Cd Co
Cadmium phosphate Cobalt phosphate
Cr
Chromium phosphate (green) Chromium phosphate (purple) Cornetite Libenthenite Torbernite Metatorbernite Nissonite Pseudomalachite Tagilite Pyromorphite Turquois Chalcosiderite Veszelyite Liberthenite
Cu
Hg Mn Ni
Mercuric phosphate Manganeous phosphate Nickel phosphate
Pb
Fluropyromorphite Chloropyromorphite Hydroxypyromorphite Pyromorphite Plumbogummite Tsumebite Dumontite Corkite Hinsdalite Parsonsite Dewindtite Renardite
Zn
Przhevalskite Embriyite Hopeite Parahopetite Phosphophyllite Scholzite Spencerite Turbuttite Zinc rockbridgeite Zinc pyromorphite Fausite
Chemical formula Cd3(PO4)2 Co3(PO4)2 Co(UO2)2(PO4)2ⴢ7H2O CrPO4 CrPO4 Cu3(PO4)(OH)3 Cu2(PO4)OH Cu(UO2)2(PO4)2ⴢ10H2O Cu(UO2)2(PO4)2ⴢ8H2O CuMg(PO4)(OH)ⴢ2.5H2O Cu5(PO4)4(OH)4 Cu2(PO4)(OH)ⴢH2O Cu5(PO4)3OH CuAl6(PO4)4(OH)8ⴢ4H2O CuFe6(PO4)4(OH)8ⴢ4H2O CuZn2PO4(OH)3ⴢ2H2O Cu2PO4OH Cu3(PO4)2 Cu3(PO4)2ⴢ3H2O Cu5O2(PO4)2ⴢ6H2O Hg2HPO4 MnPO4ⴢ1.5H2O Ni3(PO4)2 Ni(UO2)2(PO4)2ⴢ7H2O Pb10(PO4)6(F)2 Pb10(PO4)6(Cl)2 Pb10(PO4)6(OH)2 Pb3(PO4)2 PbAl3(PO4)2(OH)5ⴢH2O Pb2Cu(PO4)2(OH)3ⴢ3H2O Pb2(UO2)4(PO4)2(OH)4 PbFe3(PO4)(OH)6SO4 PbAl3(PO4)(OH)6SO4 Pb2UO2(PO4)2ⴢ2H2O Pb(UO2)4(PO4)2(OH)4ⴢ8H2O Pb(UO2)4(PO4)2(OH)4ⴢ7H2O Pb(UO2)4(PO4)2ⴢ4H2O Pb5(Cr2O4)2(PO4)2ⴢH2O Zn3(PO4)2 Zn3(PO4)2ⴢ4H2O Zn2Fe(PO4)2ⴢ4H2O Zn2Ca(PO4)2ⴢ4H2O Zn4(PO4)2(OH)2ⴢ3H2O Zn2(PO4)(OH) ZnFe4(PO4)3(OH)5 Zn5(PO4)3OH ZnAl6(PO4)4(OH)8ⴢ4H2O
Sources: Stumm and Morgan (1995); Nriagu (1984); Lindsay (1971).
Log Ksp −38.1 −4.36 −9.9 −3.07 −2.55 −48.0 −28.0 −41.0 −41.3 −23.6 −75.8 −27.9 −65.6 −179.0 −205.7 −45.8 −2.21 −3.98 −8.53 −5.20 −34.82 −8.82 −9.5 −76.8 −25.75 −18.15 −4.43 −99.3 −51.3 −91.4 −112.6 −99.1 −45.8 −92.6 −93.7 −47.4 −35.40 −17.06 −35.8 −34.1 −52.8 −26.6 −138.6 −63.1 −177.7
26
N.S. Bolan, D.C. Adriano, and R. Naidu
and variety of substitutions can occur among similar ions, rendering it almost impossible to give satisfactory unique formulae for many of the phosphate minerals. For example, mutual substitution of P for As is very common. The ability of apatite to immobilize dissolved Pb or Pb in contaminated soils through precipitation as Pb phosphate minerals has been well documented. Such precipitates are more commonly manifested as hydroxypyromorphite or as chloropyromorphite (Table 9). Two processes for the reaction of dissolved Pb with apatite have been proposed (Laperche and Traina 1998). First, Pb can react with apatite through hydroxyapatite [HA, Ca10(PO4)6(OH)2] dissolution (Eq. 6), followed by precipitation (Eq. 7) of pure hydroxypyromorphite [Pb10(PO4)6(OH)2] as described by Ma et al. (1993). Second, Pb can substitute for Ca in apatite (Eq. 8). Thus (Ca, Pb) apatite could be potentially formed by adsorption of Pb or by dissolution of HA followed by coprecipitation of mixed apatites. However, using a range of spectroscopic techniques (XRD, IR, EXAFS, and SEM), Laperche and Traina (1998) showed that, even at low Pb2+ concentrations, attenuation of aqueous Pb2+ by apatite resulted from the precipitation of pyromorphite and there was no evidence for coprecipitation as (Pb, Ca) apatite. Thus, dissolution of apatite is an important initial step in the immobilization of Pb as pyromorphite. For example, limited dissolution of PRs in alkaline soils has been shown to restrict the formation of pyromorphite (Laperche et al. 1997; Zhang et al. 1997). Ca10(PO4)6(OH)2(s) + 14H+(aq) → 10Ca2+(aq) + 6H2PO−4(aq) + 2H2O (6) 10Pb2+(aq) + 6H2PO−4(aq) + 2H2O → Pb10(PO4)6(OH)2(s) Ca10(PO4)6(OH)2(s) + xPb → (Ca10 − x) Pbx) (PO4)6(OH)2(s) + xCa 2+
(7) 2+
(8)
Two important implications of Pb–phosphate interactions deserve to be noted (Nriagu 1984). (1) The fixation of Pb2+ as insoluble phosphates in soils regulates the quantity of Pb that cycles annually in ecosystems. Formation of Pb phosphates may be one of the buffer mechanisms regulating the concentration of Pb in natural waters. (2) Pb bound to phosphate is unavailable to plants. This interaction has the beneficial effect of reducing the dietary intake of Pb by humans and herbivorous animals, as demonstrated by PBET, PBASE, and GI tests (see Table 4). For a given phosphate mineral composition, the usual stability sequence is Pb > Cu > Zn. Cadmium and Ni phosphates are remarkably sparse in the geologic record. The ionic sizes of these two metals suggest that Cd2+ and Ni2+ should be able to form several phosphates with composition analogous to those of Ca2+ and Zn2+, respectively. The dearth of these phosphates under natural conditions is related to low ionic activities of these two elements in subaqueous, supergene environments. However, many Ca and Al phosphates show significant concentrations of Cd and Ni, probably resulting from coprecipitation. Another dominant process in the adsorption of Cd2+ and Zn2+ by hydroxyapatite is via coprecipitation (Eq. 4) (Xu et al. 1994). Additionally, Soon (1981) examined the effect on the solubility of Cd in soils amended with sludges that
Phosphorus (Im)mobilization of Heavy Metals
27
had been treated with Ca(OH)2, Al2(SO4)3, or FeCl3 to precipitate P from effluent water. The sludge varied in lime equivalent and P content. At low levels of Cd addition, the solubility of Cd was controlled by adsorption, which increased with increasing pH resulting from the sludge addition. However, at high levels of Cd addition, there was evidence for the precipitation of Cd as Cd3(PO4)2 and CdCO3, which controlled the solubility. Similarly, Seaman et al. (2001) observed that hydroxyapatite addition lowered the solubility of several contaminants, such as U, Ni, Cd, Pb, and Co, in two highly weathered contaminated pond sediments, again demonstrating the potential value of hydroxyapatite for remediating metal-contaminated sites. The decrease in the solubility of these metals in the presence of hydroxyapatite was attributed to the formation of secondary metal phosphate precipitates rather than direct metal adsorption by weathered apatite grains. McGowen et al. (2001) examined the immobilization of As, Cd, Pb, and Zn in a smelter-contaminated soil using DAP. Application of high levels of DAP at a rate of 2300 mg P kg−1 was very effective in immobilizing Cd, Pb, and Zn in the contaminated soil. Activity ratio diagrams indicated that the DAP decreased solution concentration of these metals by forming metal–phosphate precipitates with low solubility products. Others have also shown that Cd3(PO4)2 can control Cd solubility in P-sufficient soils or soils amended with P (Santillan-Medrano and Jurinak 1975; Street et al. 1978). Because the solubility of Cd3(PO4)2 is too high to control the concentration of Cd in suspensions involving Fe and Al oxides and soils, it is doubtful this solid phase can play significant role in the immobilization of Cd and possibly Zn (Bolland et al. 1977; Soon 1981). Krishnamurti et al. (1996) observed that, compared with bulk soils, solidphase speciation of Cd differs substantially from root zone soils treated with P fertilizer. The amounts of Cd that was adsorbed and complexed with organic substances in the rhizosphere soil were appreciably higher than those of the corresponding bulk soil. The increase in the immobilization was attributed to precipitation as carbonate facilitated by the release of CO2 via root respiration and the organic acids released with root exudates present in the soil–root interface. It is important to emphasize that both H2PO−4 and metal ions are readily adsorbed by many arable soils, thereby maintaining typically very low concentrations of these ions in soil solution. Under this condition, precipitation of metal phosphates in soils is unlikely to occur. Nevertheless, when the H2PO−4- and metal ion-adsorbing components such as ferromanganese oxyhydroxides undergo dissolution under reducing and acidic conditions, both the metal and H2PO−4 ions are released, elevating their concentrations and subsequently causing precipitation of metal phosphates. Thus, the synergistic effects of the H2PO−4 and metaladsorbing sediments promote the formation of metal phosphates in a wide range of geologic environments. Through Liming Action of Phosphate Rocks In addition to the formation of new solid phases previously discussed, metal immobilization may also ensue
28
N.S. Bolan, D.C. Adriano, and R. Naidu
from the formation of metal oxyhydroxides due to the buffering capacity (or liming action) of hydroxyapatite (Basta et al. 2001; Chen et al. 1997b; Ma and Rao 1997). Unlike soluble P fertilizers such as SSP, TSP, and DAP, PRs neutralize acids during the dissolution reactions in soils and thus can also have some liming effect through two processes. First, most PRs contain some free calcium carbonate (CaCO3), which can act as a liming agent itself. Second, the dissolution of the apatite mineral in soils consumes acids, thereby reducing soil acidity. It is estimated that every 1 kg of P dissolved from the PR gives a liming value equivalent to 3.2 kg CaCO3. From the amounts of P and free CaCO3 present in the PR, it may be possible to calculate its equivalent total liming value. For example, a tonne of NCPR containing 13.1% P and 11.7% free CaCO3 can have a potential liming value of 536 kg CaCO3 (132 kg free CaCO3 plus 3.2 × 132 = 419 kg CaCO3 from dissolution). The equivalent liming values of various PRs range from 450 to 560 kg CaCO3 per tonne of PR, with the most liming value coming from the dissolution of apatite mineral (Table 10). Because the free CaCO3 in PRs dissolves reasonably fast, it can immediately give small amounts of liming value. On the other hand, the apatite mineral dissolves at a generally slower rate in soils, hence its
Table 10. Theoretical liming value and acidity equivalent of phosphate compounds.
Phosphate compound Diammonium phosphate Monoammonium phosphate Single superphosphate Triple superphosphate North Carolina phosphate rock Sechura phosphate rock Gafsa phosphate rock Chatham rise phosphorite Arad phosphate rock Youssafia phosphate rock Khourigba phosphate rock Egyptian phosphate rock Jordan phosphate rock Nauru phosphate rock Christmas Island Duchess phosphate rock
Total P (% w/w)
Free CaCO3 (% w/w)
Liming value (kg CaCO3 Mg−1)
Acidity equivalenta (kg CaCO3 Mg−1)
20 21 10 18
— — — —
— — — —
740 550 80 150
13.1 13.1 13.4 8.9 14.1 13.8 14.4 13.0 13.4 15.6 16.4 13.5
11.7 5.1 7.1 27.6 8.2 5.4 6.1 4.9 7.7 4.1 2.1 1.8
536 470 500 560 533 495 520 465 505 540 545 450
— — — — — — — — — — — —
aEquivalent acidity is kg calcium carbonate required to neutralize the acidity from 1000 kg (1 Mg) of fertilizer.
Phosphorus (Im)mobilization of Heavy Metals
29
liming value persists over a longer period of time. It is to be noted that some of the unreactive PRs such as Christmas Island PR, Nauru PR, and Duchess PR can also have significant amount of liming value. However, because these unreactive PRs are unlikely to dissolve in soils, no benefit from adding such PRs either as a P source or as a liming material should be expected. The potential value of PRs as a liming material in mitigating acid mine drainage has been documented both under laboratory and field conditions (Frazer 2001). Treating acid mine drainage with PRs not only neutralizes the acid through their buffering action but also reduces the solution concentration of metals through precipitation and ion-exchange reactions (Feng et al. 2000). E. Rhizosphere Modifications P compounds affect metal transformation in the rhizosphere primarily through their effects on soil pH (i.e., acidification) and mycorrhizal association. Acidification pH is one of the most important factors that control the reactions of metals in soils, and acidification of soils has often been shown to enhance the mobilization of metals (Adriano 2001). Fertilizer use in managed ecosystems is one of the major contributors to soil acidification; thus, it is important to understand the acidifying effect of P fertilizer and its potential ecological consequences before any decision can be made on the choice of the P fertilizer in relation to immobilization of metals. The acidifying effect of fertilizers is quantified as ‘acidity equivalent,’ which gives the amount of pure CaCO3 required to neutralize the acid produced by a given weight of the fertilizers (see Table 10). Superphosphates and ammonium phosphates are the most common soluble P fertilizers used in agriculture. The dissolution of MCP in soils results in the formation of DCP (Eq. 1) with a release of phosphoric acid close to the fertilizer granules. Nitrogen-containing P fertilizers such as MAP and DAP cause soil acidification in two ways. First, the natural process of nitrification, where NH+4 ions are converted to NO−3 ions, results in the release of H+ ions (Eq. 2). Second, NO−3 ions, not being strongly adsorbed by the soil, induce the leaching of basic cations to maintain the charge balance. It is the depletion of basic cations during the leaching of NO−3 ions that accelerates the acidification process and the drop in pH. In legume-based pastures and crop production systems, P fertilizer is added mainly to promote N fixation by the legumes by overcoming soil P deficiency. Thus, it is important to point out that irrespective of the P fertilizer source, application of P to legume-based systems promotes N fixation, thereby indirectly causing soil acidification. The amount of acidity produced indirectly by N fixation is mainly influenced by the extent of NO−3 leaching, which could be much higher than the acidity produced directly by the dissolution of MCP in superphosphate fertilizer. Rhizosphere acidification is considered to be one of the main reasons for the accumulation of metals by legumes (Marschner 1995). Acidification caused by P fertilizers induces the mobilization of metals. For example, application of DAP has been shown to enhance metal solubilization
30
N.S. Bolan, D.C. Adriano, and R. Naidu
through acidification (McGowen et al. 2001). The subsequent effect on the uptake of metals depends on the relative amount of NH+4 uptake. To maintain the electrical charge inside the plants, plants using the positively charged NH+4 ions as the main source of N tend to accumulate fewer metal cations (Andrew and Johnson 1976; Soon and Miller 1977). Similarly, Hettiarachchi et al. (2000) have observed that preacidification of Pb-contaminated soils enhances the solubilization of Pb and the subsequent precipitation by phosphate during the in situ stabilization of Pb using P fertilizer. Acidification affects the transformation of metal ions through several ways, foremost of which include (a) modification of surface charge in variable charge soils; (b) altering the speciation of metals; and (c) influencing the reduction and oxidation reactions of the metals (Adriano 2001). Adsorption of metals almost invariably decreases with increasing soil acidity (or decreasing pH) (Basta and Tabatabai 1992; Bolan et al. 1999a; Naidu et al. 1994; Tiller 1988). Three possible reasons have been advanced for this phenomenon (Naidu et al. 1994). First, in variable charge soils, a decrease in pH causes a decrease in surface negative charge, resulting in lower cation adsorption. Second, a decrease in soil pH is likely to decrease hydroxy species of metal cations (MOHn+), which are adsorbed preferentially over mere metal cation (Hodgson et al. 1964). For example, Naidu et al. (1994) observed that CdOH+ species which dominate above pH 8 have greater affinity for adsorption sites than just Cd2+. Third, acidification causes the dissolution of metal compounds, resulting in an increase in the concentration of metals in soil solution. In general, Cd uptake by plants increases with decreasing pH. For example, higher Cd uptake was obtained for lettuce and Swiss chard grown on acid soils (pH 4.8–5.7) than on calcareous soils (pH 7.4–7.8) (Mahler et al. 1978). Consequently, it is recommended that soil pH be maintained at pH 6.5 or greater in lands receiving biosolids containing Cd. Acidification affects the leaching and residence time of many trace elements. Taylor (1975) observed that the amount of trace elements released from the mor soils of Sweden increased with decreasing pH of the acid precipitation. Approximately 85% of the total Cd was released from the soil at pH 2.8. The time needed for a 10% decrease in the total concentration of Cd in the mor horizon through leaching was estimated to be 1.7 yr at pH 2.8, 4–5 yr at pH 3.2, and 20 yr at pH 4.2. In the case of metalloids, such as As, the effect of soil acidity on adsorption is manifested through two interacting factors: the increasing negative surface potential on the plane of adsorption and the increasing amount of negatively charged As(V) species present in soil solution. Although the first factor results in lower As(V) adsorption, the second factor is likely to increase adsorption. Thus, the pH effect on As(V) adsorption is largely influenced by the nature of the mineral surface. For example, in soils with low oxide content, increasing pH had little effect on adsorption, whereas in highly oxidic soils, adsorption decreased with increasing pH (Smith et al. 1999). In general, adsorption of
Phosphorus (Im)mobilization of Heavy Metals
31
As(V) decreases with increasing pH. However, in contrast to As(V), adsorption of As(III) tends to increase with increasing pH (Smith et al. 1998). Soil acidification affects the solubility of Cr through its effect on adsorption/ precipitation and oxidation/reduction reactions (Bartlett 1991; James 1996). The adsorption of Cr(VI) in soil increases with decreasing pH, but the adsorption of Cr(III) decreases (Bartlett and Kimble 1976). Similarly, while the reduction of Cr(VI) to Cr(III) (a H+ consumption reaction, Eq. 9) increases with decreasing soil pH (James 1996), the oxidation of Cr(III) to Cr(VI) (a H+ donation reaction) decreases (Bartlett and James 1979). The reduction of Cr(VI) to Cr(III) occurs readily in most soils, especially arable soils, due to the presence of organic matter (Eq. 3), whereas the oxidation of Cr(III) to Cr(VI) requires the presence of oxidized Mn in the soil as an electron acceptor for the reaction to proceed (Barlett and James 1979). 2Cr2O7 + 3C0 + 16H+ → 4Cr3+ + 3CO2 + 8H2O
(9)
In general, with the exception of Se and Mo, trace elements are more soluble in soils at low pH because of the dissolution of the carbonates, phosphates, and other solid phases. Low pH also lowers the CEC of organic matter and mineral surfaces, thereby weakening the sorption of metals to specific adsorption sites; this explains why the effect of hydroxyapatite in the immobilization of metals is enhanced through preacidification of the contaminated soils (Hettiarachchi et al. 2000). Mycorrhizal Association Mycorrhizal association is considered to be one of the main factors responsible for metal tolerance in plants and is likely to have a major influence on the hyperaccumulation of metals by certain plant species (Khan et al. 2000). The potential of phytoremediation of metal-contaminated sites can be enhanced by inoculating hyperaccumulating plants with mycorrhizal fungi that are most appropriate for contaminated sites. The effect of mycorrhizae on P-induced (im)mobilization of metals is mediated through two processes. First, mycorrhizae have often been shown to increase the uptake of P by plants through greater exploration of the bulk soil and enhanced solubilization of water-insoluble P compounds, such as PRs (Bolan 1991). Enhanced solubilization of P compounds by the mycorrhizal fungi can induce mobilization and the subsequent uptake of metals associated with the P compounds. On the other hand, it can also facilitate the release of P from water-insoluble P compounds for the immobilization of metals in the soils. Second, a high level of P in soils has been shown to inhibit mycorrhizal infection (Amijee et al. 1989), which is likely to affect the mobilization of metals by the fungi. The evidence for these processes is discussed next. The effectiveness of mycorrhizal association on the growth and P uptake from slow-release P compounds, such as PRs with different solubilities, has been elucidated (Table 11). The premise has been that although mycorrhizal plants are better than nonmycorrhizal plants in extracting P from PRs, the in-
32
N.S. Bolan, D.C. Adriano, and R. Naidu
Table 11. The effectiveness of water-insoluble phosphate fertilizers in mycorrhizal versus nonmycorrhizal plants. Plant species Bouteloua (Bouteloua gracilis) Lavender (Lavendula spica) Tomato (Lycopersicum escelenium)
Oil palm (Elaeis sp.) Leucaena (Leucaena leucocephala) Stylosanthes (Stylosanthes guyanensis) Clover (Trifolium pratense) Centrosems (Centrosems sp.) Stylosanthes (Stylosanthes guyanensis) Sorghum (Sorghum vulgare) American elms (Ulmus americana) Soybean (Glycine max)
Pueraria (Pueraria sp.) Stylosanthes (Stylosanthes sp.) Subclover (Trifolium subterraneum) Ryegrass (Lolium perenne) Subclover (Trifolium subterraneum) White clover (Trifolium repens) White clover (Trifolium repens) Ryegrass (Lolium perenne) Lucerne (Medicago sativa)
Phosphate compound
Effectivenessa
Calcium phytate Phosphate rock Dicalcium phosphate Tricalcium phosphate Phosphate rock Phosphate rock Phosphate rock Phosphate rock Hyperphosphate Phosphate rock Phosphate rock Phosphate rock Phosphate rock Tricalcium phosphate Phosphate rock Aluminium phosphate Iron phosphate Phosphate rock Phosphate rock Phosphate rock Phosphate rock Phosphate rock Iron phosphate Iron phosphate Iron phosphate Iron phosphate Calciphos Phosphate rock Phosphate rock Phosphate rock Phosphate rock Tricalcium phosphate
1.01 5.00 1.56 5.38 1.96 2.20 4.00 1.32 1.69 5.00 3.01 1.39 2.38 2.80 5.52 1.78 5.35 2.11 1.45 9.95 1.11 1.52 1.85 5.84 11.19 3.43 1.49 1.45 1.51 3.75 3.01 1.52
aCalculated from the ratio of yields of mycorrhizal to nonmycorrhizal plants at a particular level of P or from the ratio of fertilizer P required for nonmycorrhizal to mycorrhizal plants to produce 50% of maximum yield. Values >1 indicate greater effectiveness of mycorrhizal compared to nonmycorrhizal plants. Source: Bolan (1991).
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crease in P uptake from PRs is identical to that from soluble P fertilizers. Similarly, the increase in the availability of P for mycorrhizal plants from P that had been allowed to react with soil is similar to that from freshly applied P. These studies were not able to provide conclusive evidence for the dissolution of water-insoluble fertilizers by mycorrhizae (Ness and Vlek 2000). Somewhat contradicting, Bolan (1991) noticed that the increase in P uptake with mycorrhizal infection varied with the P solubility, with greatest benefit arising from the least soluble P source. The logical explanation for the increased uptake by mycorrhizal plants in this case is greater surface volume in the bulk soil explored by the fungi. A further possibility is that mycorrhizal hyphae may be able to chemically modify the availability of less-soluble P sources by producing organic compounds with chelating properties, such as citrate. In addition, production of phosphatases by ectomycorrhizal fungi is important in the solubilization of organic phytates, which constitute a large fraction of total P in humic soils (Bartlett and Lewis 1973; Mitchell and Read 1981; Williamson and Alexander 1975). Similarly, ectomycorrhizae have been shown to produce large amounts of calcium oxalate (Lapeyrie 1988; Malajczuk and Cromack 1982), which may be involved in the chelation of Fe and Al, thereby releasing P for plant uptake (Graustein et al. 1977; Treeby et al. 1989). Some experimental evidence for direct chemical modification of P availability by endomycorrhizal plants is available (Abbott and Robson 1982; Gianinazzi-Pearson and Gianinazzi 1978). Parfitt (1979) suggested that the increased uptake of P from goethite–phosphate complexes by mycorrhizal plants might result from increased production of citrate and other organic compounds. Similarly, Jayachandran et al. (1989) have observed that in the presence of synthetic chelates (EDDHA) mycorrhizae caused greater uptake of P than in the absence of these chelates, whereas nonmycorrhizal plants were unable to take advantage of the P released by chelation. In this case, siderophore production by mycorrhizal fungi or other soil microbes could significantly increase P availability in lowpH soils, and this is a feasible mechanism by which mycorrhizal plants could acquire P sources unavailable to nonmycorrhizal plants. Differences in the absorption of anions and cations by mycorrhizal and nonmycorrhizal plants may lead to differences in rhizosphere pH (Buwalda et al. 1983), which may then affect the availability of adsorbed P and metals to plants (Hedley et al. 1982). Because mycorrhizal plants utilize NH+4−N more efficiently than nonmycorrhizal plants (Smith et al. 1985), this could create pH differences in the rhizosphere. It is logical that H+ extrusion, which is an inevitable consequence of NH+4−N assimilation in cells (Bolan et al. 1991; Raven and Smith 1976), would occur from the hyphae as well as from the roots (Raven et al. 1978). This extrusion could reduce the pH around the infected root, thereby enhancing the availability of P and the associated metals from sparingly soluble P sources such as PRs. Colonization of plant roots with mycorrhizal fungi could have synergistic or antagonistic effects on the rhizosphere microflora (Ames et al. 1984; Meyer and Linderman 1986). Consequently the changes in rhizosphere microflora indirectly
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affect the availability of both organic and inorganic P sources and metals to plants (Azcon et al. 1976). The effect of P on mycorrhizal infection and the subsequent uptake of trace elements by plants has been explored (Gildon 1983; Gildon and Tinker 1983; Lambert et al. 1979; Pacovsky 1986; Timmer and Leydon 1980). Findings indicate that high levels of bioavailable P in soils inhibited mycorrhizal infection, concomitantly reducing the uptake of metals. For example, Lambert et al. (1979) observed that high levels of P in soils reduced mycorrhizal infection, lowering Zn uptake. Similarly, Pacovsky (1986) observed that mycorrhizal plants increased the uptake of Zn and Cu but this decreased with increasing level of soil P. In examining the relationship of mycorrhizal infection to P-induced Cu deficiency in sour orange seedlings, Timmer and Leydon (1980) concluded that, in the case of nonmycorrhizal plants, P induces Cu deficiency by stimulating growth of plants until Cu becomes a limiting nutrient, whereas in the case of mycorrhizal plants Cu deficiency occurs because of direct P-induced inhibition of mycorrhizal development. The antagonistic effect of P on mycorrhizal infection is likely to have a strong influence on phytoremediation of metals in contaminated soils. There has been some indication that mycorrhizal infection, particularly with the heavy metal tolerant isolate, could protect plants against heavy metal phytotoxicity (Gildon and Tinker 1981; Griffioen 1994; Griffioen et al. 1994). These infections might be important in revegetation of polluted sites, and any attempt to immobilize metals using high levels of soluble P compounds is likely to affect the mycorrhizae, leading to poor establishment of plant colonies.
Summary A large number of studies have provided conclusive evidence for the potential value of both water-soluble (e.g., DAP) and water-insoluble (e.g., apatite, also known as PRs) P compounds to immobilize metals in soils, thereby reducing their bioavailability for plant uptake. It is, however, important to recognize that, depending on the nature of P compounds and the heavy metal species, application of these materials can cause either mobilization or immobilization of the metals. Furthermore, some of these materials contain high levels of metals and can act as an agent of metal introduction to soils. Accordingly, these materials should be scrutinized before their large-scale use as immobilizing agent in contaminated sites. Although mobilization by certain P compounds enhances the bioavailability of metals, immobilization inhibits their plant uptake and reduces their transport in soils and subsequent groundwater contamination. Whenever phytoremediation of contaminated sites is practicable, appropriate P compounds can be used to enhance the bioavailability of metals for plant uptake. Removal of metals through phytoremediation techniques and the subsequent recovery of the metals or their safe disposal are attracting research and commercial interests. Phosphate
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compounds can be used to enhance the solubilization of metals, leading to their increased uptake by plants. However, when it is not possible to remove the metals from the contaminated sites by phytoremediation, other viable options such as in situ immobilization should be considered as an integral part of risk management. One way to facilitate such immobilization is by altering the physicochemical properties of the metal–soil complex by introducing a multipurpose anion, such as phosphate, that enhances metal adsorption via anion-induced negative charge (i.e., CEC) and metal precipitation. It is important to recognize that large-scale use of P compounds can lead to surface and groundwater contamination of this element. It is therefore, important that future research should aim to focus on the role of P compounds on in situ remediation and natural attenuation in metal-contaminated sites, with minimum impact of P on quality of water sources.
Acknowledgments The U.S. Department of Energy contract number DE-FC-09–96SR18546 with the University of Georgia’s Savannah River Ecology Laboratory supported writing and editing time for Drs. Bolan and Adriano.
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Qafoku NP, Kukier U, Sumner ME, Miller WP, Radcliffe DE (1999) Arsenate displacement from fly ash in amended soils. Water Air Soil Pollut 114:185–198. Qafoku NP, Dudka S, Sumner ME, Miller WP (2001) Arsenic, boron, selenium, and molybdenum displacement and transport in a fly ash amended soil leached with calcium phosphate solution. Commun Soil Sci Plant Anal 32:1499–1512. Rai D, Earty LE, Zachara JM (1987) Chromium(III) hydrolysis constant and solubility of chromium(III) hydroxide. Inorg Chem 26:345–349. Rajan SSS (1978) Sulfate adsorbed on hydrous alumna, ligands displaced, and changes in surface charge. Soil Sci Soc Am J 42:39–44. Rajan SSS, Perrott KW, Saunders WMH (1974) Identification of phosphate-reactive sites of hydrous alumna from proton consumption during phosphate adsorption at constant pH values. J Soil Sci 25:438–447. Rajan SSS, Watkinson JH, Sinclair AG (1996) Phosphate rocks for direct application to soils. Adv Agron 57:77–159. Raven JA, Smith FA (1976) Nitrogen assimilation and transport in vascular land plants in relation to intercellular pH regulation. New Phytol 76:415–431. Raven JA, Smith SE, Smith FA (1978) Ammonium assimilation and the role of mycorrhizas in climax communities in Scotland. Trans Bot Soc Edinb 43:27–35. Reynolds JG, Naylor DV, Fendorf SE (1999) Arsenic sorption in phosphate-amended soils during flooding and subsequent aeration. Soil Sci Soc Am J 63:1149–1156. Roberts A, Longhurst RD, Brown MW (1994) Cadmium status of soils, plant and grazing animals in New Zealand. N Z J Agric Res 33:119–129. Ross SM (1994) Retention, transformation and mobility of toxic metals in soils. In: Ross SM (ed) Toxic Metals in Soil–Plant Systems. Wiley, New York, pp 63–152. Rothbaum HP, Goguel RL, Johnson AE, Mattingly GEG (1986) Cadmium accumulation in soils from long continued application of superphosphate. J Soil Sci 37:99–107. Ruby MV, Davis A, Schoof R, Eberle S, Sellstone CM (1994) In situ formation of lead phosphates in soils as a method to immobilise lead. Environ Sci Technol 28:646–654. Ruby MV, Davis A, Schoof R, Eberle S, Sellstone CM (1996) Estimation of bioavailability using a physiologically based extraction test. Environ Sci Technol 30:420–430. Rumberg CB, Engel RE, Meggitt WF (1960) Effect of phosphorus concentration on the absorption of arsenate by oats from nutrient solution. Agron J 52:452–453. Ryden JC, Syers JK (1976) Calcium retention in response to phosphate adsorption by soils. Soil Sci Soc Am J 40:845–846. Santillan-Medrano J, Jurinak JJ (1975) The chemistry of lead and cadmium in soil: solid phase formation. Soil Sci Soc Am Proc 39:851–856. Sauve S, Norvell WA, McBride M, Hendershot W (2000) Speciation and complexation of cadmium in extracted soil solutions. Environ Sci Technol 34:291–296. Sawhney BL (1974) Charge characteristics of soils as affected by phosphate sorption. Soil Sci Soc Am J 38:159–160. Schalscha EB, Pratt PF, Soto D (1974) Effect of phosphate adsorption on the cationexchange capacity of volcanic ash soils. Soil Sci Soc Am J 38:539–540. Schippers A, Sand W (1998) Bacterial leaching of metal sulfides proceeds by two indirect mechanisms via thiosulfate or via polysulfides and sulfur. Appl Environ Microbiol 65:319–321. Schwab AP (1989) Manganese-phosphate solubility relationships in an acid soil. Soil Sci Soc Am J 53:1654–1660. Seaman JC, Arey JS, Bertsch PM (2001) Immobilization of nickel and other metals in contaminated sediments by hydroxyapatite addition. J Environ Qual 30:460–469.
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Shuman LM (1986) Effect of ionic strength and anions on zinc adsorption by two soils. Soil Sci Soc Am J 50:1438–1442. Shuman LM (1991) Chemical forms of micronutrients in soils. In: Mortvedt JJ, Cox FR, Shuman LM, Welch RM (eds) Micronutrients in Agriculture. Soil Science Society of America, Madison, WI, pp 113–144. Singh BR, Myhr K (1998) Cadmium uptake by barley as affected by Cd sources and pH levels. Geoderma 84:185–194. Smith SE, St John BJ, Smith FA, Nicholas DJD (1985) Activity of glutamine synthetase and glutamine dehydrogenase in Trifolium subterraneum L. and Allium cepa L.: effect of mycorrhizal infection and phosphate nutrition. New Phytol 99:211–217. Smith E, Naidu R, Alston AM (1998) Arsenic in the soil environment: a review. Adv Agron 66:149–195. Smith E, Naidu R, Alston AM (1999) Chemistry of arsenic in soils. I. Sorption of arsenate and arsenite by four Australian soils. J Environ Qual 28:1719–1726. Smith E, Naidu R, Alston A (2002) Chemistry of arsenic in soils. II. Effect of pH and ionic strength. J Environ Qual 31:557–563. Soon YK (1981) Solubility and sorption of cadmium in soils amended with sewage sludge. J Soil Sci 32:85–95. Soon YK, Miller MH (1977) Changes in the rhizosphere due to NH+4 and NO−3 fertilization and phosphorus uptake by corn seedlings (Zea mays L.). Soil Sci Soc Am Proc 41:77–80. Snoeyink VL, Jenkins D (1980) Water Chemistry, Wiley, New York. Sparrow LA, Salardini AA, Bishop AC (1993) Field studies of cadmium in potatoes (Solanum tuberosum L). 1. Effects of lime and phosphorus on cv. russet burbankaust. J Agric Res 44:845–853. Sposito G (1984) The Surface Chemistry of Soils. Oxford University Press, New York. Sposito G, Lund LJ, Chang AC (1982) Trace metal chemistry in arid zone field soils amended with sewage sludge. I. Fractionation of Ni, Cu, Zn, Cd and Pb in solid phases. Soil Sci Soc Am J 46:260–264. Street JJ, Sabey BR, Lindsay WL (1978) Influence of pH, phosphorus, cadmium, sewage sludge, and incubation time on the solubility and plant uptake of cadmium. J Environ Qual 7:286–290. Stumm W, Morgan JJ (1995) Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters, 3rd Ed. Wiley, New York. Suzuki T, Hatsushika T, Hayakawa Y (1981) Synthetic hydroxyapatites employed as inorganic cation-exchangers. J Chem Soc Faraday Trans 77:1059–1062. Syers JJ, MacKay AD, Brown MW, Currie LD (1986) Chemical and physical characteristics of phosphate rock materials of ranging reactivity. J Sci Food Agric 37:1057– 1064. Taylor G (1975) Leaching rates of heavy metal ions in forest soils. Water Air Soil Pollut 9:137–148. Tessier A, Campbell PGC, Bissom M (1979) Sequential extraction procedure for the speciation of particulate trace metals. Anal Chem 51:844–850. Tiller KG (1988) Heavy metals in soils and their environmental significance. Adv Soil Sci 9:113–142. Timmer LW, Leydon RF (1980) The relationship of mycorrhizal infection to phosphorusinduced copper deficiency in sour orange seedlings. New Phytol 85:15–23. Treeby M, Marschner H, Ro¯mheld V (1989) Mobilization of iron and other micronutrient
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cations from a calcareous soil by plant-borne, microbial and synthetic metal chelators. Plant Soil 114:217–226. Tu C, Zheng CR, Chen HM (2000) Effect of applying chemical fertilizers on forms of lead and cadmium in red soil. Chemosphere 41:133–138. Wann SS, Uehara G (1978a) Surface charge manipulation in constant surface potential soil colloids. I. Relation to sorbed phosphorus. Soil Sci Soc Am J 42:565–570. Wann SS, Uehara G (1978b) Surface charge manipulation in constant surface potential soil colloids. II. Effect on solute transport. Soil Sci Soc Am J 42:886–888. Whitelaw MA (2002) Growth promotion of plants innoculated with phosphate-solubilizing fungi. Adv Agron 69:99–151. Williams CH, David DJ (1976) The accumulation in soil of cadmium residues from phosphate fertilizers and their effect on the cadmium content of plants. Soil Sci 121: 86–93. Williamson B, Alexander I (1975) Acid phosphatases localized in the sheath of beech mycorrhizas. Soil Biol Biochem 7:194–198. Woolson EA, Axley JH, Kearney P (1973) The chemistry and phytotoxicity of arsenic in soil: II. Effect of time and phosphorus. Soil Sci Soc Am Proc 37:254–258. Xie RJ, Mackenzie AF (1991) Molybdate sorption-desorption in soils treated with phosphate. Geoderma 48:321–333. Xie RJ, Mackenzie AF, Lou ZJ (1993) Causal-modeling pH and phosphate effects on molybdate sorption in 3 temperate soils. Soil Sci 155:385–397. Xu Y, Schwartz FW, Traina SJ (1994) Sorption of Zn2+ and Cd2+ on hydroxyapatite surfaces. Environ Sci Technol 28:1472–1480. Zhang P, Ryan JA, Yang J (1997) In vitro soil Pb solubility in the presence of hydroxyapatite. Environ Sci Technol 32:2763–2768. Manuscript received February 16; Accepted March 7, 2002.
Springer-Verlag 2003
Rev Environ Contam Toxicol 177:45–122
Environmental Fate of Methyl Bromide as a Soil Fumigant Scott R. Yates, Jay Gan, and Sharon K. Papiernik Contents I. Introduction ....................................................................................................... 46 A. Environmental Concerns .............................................................................. 46 B. Economic Concerns ..................................................................................... 47 II. Chemical and Physical Properties of Methyl Bromide ................................... 49 III. Methyl Bromide Use as a Soil Fumigant ........................................................ 50 A. History and Scope of Use ........................................................................... 50 B. Application Methods .................................................................................... 51 IV. Ozone Depletion and Methyl Bromide ............................................................ 52 A. Reactions with Ozone .................................................................................. 52 V. Sampling and Analysis of Methyl Bromide in the Air ................................... 54 A. Container Methods ....................................................................................... 56 B. Adsorbent Methods ...................................................................................... 56 C. Other Methods .............................................................................................. 58 VI. Processes Affecting Environmental Fate of Methyl Bromide ........................ 58 A. Transformation ............................................................................................. 58 B. Phase Partitioning ........................................................................................ 68 VII. Simulating the Environmental Fate of Methyl Bromide ................................. 70 A. Transport Model .......................................................................................... 70 B. Mobility Indices ........................................................................................... 73 C. Simulating Transport in Relatively Dry Soils ............................................. 74 VIII. Methyl Bromide Diffusion in Soils ................................................................. 77 A. Diffusion Coefficient ................................................................................... 77 B. Methyl Bromide Diffusion in Soils ............................................................. 77 IX. Assessing Methyl Bromide Volatilization from Soil ...................................... 81 A. Measurement Methods ................................................................................. 81 B. Field Experiments to Determine Methyl Bromide Volatilization .............. 89 X. Potential Methods for Minimizing Volatilization ............................................ 94 A. Containment ................................................................................................. 94 B. Soil Conditions ............................................................................................. 102 C. Effect of Degradation Rate on Emissions ................................................... 104
Communicated by George W. Ware. S.R. Yates ( ), S.K. Papiernik USDA-ARS, George E. Brown Jr. Salinity Laboratory, 450 West Big Springs Road, Riverside, CA 92507-4617, USA J. Gan Department of Environmental Science, University of California, 417 Geology Riverside, CA, USA 92521
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XI. Considerations for Developing Alternatives to Methyl Bromide ................... 106 Summary .................................................................................................................... 108 References .................................................................................................................. 109
I. Introduction Methyl bromide (bromomethane, MeBr) has been used widely since the 1940s as an effective preplant soil fumigant for controlling nematodes, plant pathogens, weeds and insects (UNEP 1995; Noling and Becker 1994). Fumigant use is vital for the economic viability of many crops, including strawberries, tomatoes, peppers, eggplants, tobacco, ornamentals, nursery stock, vines, and turf (Anderson and Lee-Bapty 1992; NAPIAP 1993; Ferguson and Padula 1994). The total global sales of MeBr were 7.16 × 107 kg in 1992, about 75% of which was used as a preplant soil fumigant (UNEP 1995). Methyl bromide is also widely used as a structural and commodity fumigant, as well as for quarantine or regulatory purposes (Anderson and Lee-Bapty 1992; NAPIAP 1993; Ferguson and Padula 1994; UNEP 1995). Its success as a fumigant is largely due to its wide spectrum of activity against pests at many stages of life, its ability to penetrate the fumigated zones, and the ease of application. Because of its high volatility, it leaves very low residue levels in the soil that may be phytotoxic or accumulated in plants, a problem commonly associated with the use of many other modern pesticides. A. Environmental Concerns In 1991, MeBr was identified as a potential ozone-depleting compound (Chakrabarti and Bell 1993). In 1992, on the Fourth Conference of the Parties to the Montreal Protocol, MeBr was officially added to the list of ozone-depleting chemicals, with its production suggested to be frozen at the 1991 level, effective from 1995. The inclusion of MeBr in the ozone-depleting chemicals list naturally brought this fumigant within the scope of the U.S. Environmental Protection Agency (EPA) Clean Air Act, which has an amendment that mandates discontinuation of any chemical with an ozone depletion potential (ODP) greater than 0.2 at the beginning of 2001. The ODP index for MeBr was determined to be 0.60–70 in 1992 (UNEP 1995); the ODP estimate was reduced to 0.4 in 1998. In March 1993, EPA announced that MeBr was scheduled for a phaseout in the United States by the year 2001 (USEPA 1993). This date was later changed to 2005 (USEPA 2000). During the past decade, there has been an increased research effort devoted to understanding the effects of halogenated gases emitted into the atmosphere on the depletion of the stratospheric ozone layer. According to the Ozone Assessment Synthesis Panel of the United Nations Environmental Programme (UNEP), the hole in the Antarctic ozone layer is due primarily to increases in chlorine- and bromine-containing chemicals in the atmosphere. Even though most of the ozone loss is due to chlorinated compounds (90%–95%; Watson et al. 1992), attention has been focused more recently on MeBr because strato-
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spheric bromine is believed to be 40 times more efficient than chlorine in breaking down ozone on a per atom basis (Wofsy et al. 1975). Although the largest effects from ozone-depleting gases have been observed in the southern hemisphere, there are indications that atmospheric ozone is also decreasing in the northern hemisphere. There is a great deal of uncertainty in estimates of the global MeBr budget. In the early 1990s, the ocean was viewed as a net source of MeBr. More recent global balances account for larger sinks than sources (Yvon-Lewis and Butler 1997), with the ocean acting as a net sink of MeBr, the magnitude of which is being refined (Lobert et al. 1997; King et al. 2000). Soil fumigation is thought to contribute 32 Gg yr−1 (1 Gg is equivalent to 1000 metric tons) of MeBr to the atmosphere, or ⬃20% of the total MeBr source (Yvon-Lewis and Butler 1997). The oceans represent the largest known source of atmospheric MeBr, followed by fumigation (Butler 2000). Other natural sources of atmospheric MeBr include biomass burning and production by plants, salt marshes, and fungi (Butler 2000). In recent global budgets, only 60% of the MeBr sinks were accounted for by the quantified source terms, and the “missing source” outweighed all other sources in the budget (Butler 2000). Agricultural use of MeBr, including soil fumigation, may be responsible for 3%–10% of stratospheric ozone depletion (USDA 2001). The relative significance of each global source of MeBr, including that from agricultural uses, needs to be better quantified to assist in developing rational national and worldwide policy. B. Economic Concerns An economic assessment conducted by the U.S. Department of Agriculture (USDA) indicated that the phase-out of MeBr as a fumigant will have a substantial impact on many commodities because current alternative control practices are either less effective or more expensive than MeBr (NAPIAP 1993; Ferguson and Padula 1994). A recent estimate of the annual economic loss to U.S. producers and consumers resulting from a ban of agricultural uses of MeBr is $500 million (Carpenter et al. 2000). In addition, currently a single chemical alternative that can completely replace MeBr does not exist (Anderson and Lee-Bapty 1992; Duafala 1996). Under these circumstances, MeBr has become the topic of many heated discussions, and the “methyl bromide issue” has received widespread attention (Anonymous 1994; Noling and Becker 1994; Taylor 1994; Sauvegrain 1995; Butler 1995, 1996; Duafala 1996; Thoms 1996). The many unanswered questions have also stimulated extensive research on the environmental fate of MeBr, particularly on estimating the relative contribution from each source, obtaining accurate estimates for volatilization losses from fumigated fields, understanding the processes and factors that affect the volatilization, and identifying and developing emission reduction techniques. The USDA National Agricultural Pesticide Impact Assessment Program conducted one of the first assessments of the economic impact of eliminating MeBr (NAPIAP 1993). This assessment determined that there would be a substantial
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adverse impact on the agricultural community and that this would be most strongly noticed in two states, California and Florida, the primary users of MeBr. It was estimated that a MeBr phase-out for preplant soil fumigation would cause $1.5 billion dollars in annual lost production in the U.S. This estimate ignored postharvest, nonquarantine uses, and quarantine treatments of imports and other future economic aspects such as lost jobs, markets, etc. The report predicted that the major crop losses would occur with tomatoes ($350M), ornamentals ($170M), tobacco ($130M), peppers ($130M), strawberries ($110M), and forest seedlings ($35M). More recently, the National Center for Food and Agricultural Policy (NCFAP) conducted an assessment of the economic implications of the methyl bromide ban (Carpenter et al. 2000). The NCFAP estimates a smaller economic loss of $479M to producers and consumers with the ban of preplant uses of methyl bromide. These losses are due to decreases in yield with use of alternative pest control strategies, increased production costs, and changes in the marketing window in response to supply and demand. The NCFAP estimates that losses of $235M may occur in annual crops (such as tomatoes, strawberries, and peppers), $143M in perennials (orchards and grapes), and $101M in nurseries and ornamentals. As research on methyl bromide alternatives continues to progress and regulatory issues surrounding soil fumigation take effect, the economic impact of the MeBr ban will become more clearly defined. The purpose of this review is to summarize studies on the transformation and transport processes of MeBr in soil, the interactions of these processes, and their effect on volatilization of MeBr into the atmosphere. Special emphasis is given to recent field, laboratory, and modeling studies that have been conducted for determining MeBr volatilization losses under various conditions and for identifying approaches to minimize these losses. Such a review has not been written, although a number of reviews have appeared on other aspects of MeBr use, such as the toxicological effects of MeBr (Alexeeff and Kilgore 1983; Yang et al. 1995), efficiency and phytotoxicity (Maw and Kempton 1973), and application and movement of fumigants in general (Goring 1962; Hemwall 1962; Hoffmann and Malkomes 1978; Siebering and Leistra 1978; Munnecke and Van Gundy 1979; Lembright 1990). A purpose of this review is to summarize relevant information concerning MeBr to provide a reference source to decision makers as well as to scientists in related industrial and academic areas. It also contains detailed information on measurements of volatilization on various scales (e.g., field, miniplots, and laboratory columns) that may be applicable for studies of other volatile pesticides or organic compounds in general. Reported studies on the use of MeBr in fumigating stored products, commodities, and structures are not included. Most of the MeBr studies completed in the early years are not cited because these studies dealt with issues of efficacy, rather than environmental fate. As the use of MeBr in soil fumigation has spanned more than five decades, there is a wealth of research on this topic, and many interesting research investigations have been conducted. Failure to include an
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article in this review should not be construed as a negative implication and is the result of our attempt to prepare a relatively concise review article. It is our belief that such a review is not only timely for MeBr itself but also offers lessons that will be valuable in finding and developing successful alternative fumigants to replace MeBr. Many of the factors affecting the phase-out of MeBr also arise for other chemical fumigants, most notably their potential to volatilize and contaminate the atmosphere. Fortunately, the emission control strategies discussed herein will also assist in reducing emissions of other fumigant compounds (e.g., 1,3-dichloropropene, chloropicrin, methyl isothiocyanate) and emerging potential chemical alternatives (e.g., methyl iodide, propargyl bromide).
II. Chemical and Physical Properties of Methyl Bromide Some of the basic physical and chemical properties of MeBr are listed in Table 1. Because of its high vapor pressure, MeBr can readily penetrate many matrices and is extremely difficult to contain even in the laboratory. As MeBr is colorless and odorless at room temperature, even at potentially toxic concentrations, severe exposure can occur unknowingly (Yang et al. 1995). In commercial formulations of MeBr, various percentages (0.5%–33%) of chloropicrin are added as a warning agent to protect workers and residents during and immediately after MeBr applications and to assist in protecting plants from disease. However, it Table 1. Selected physical properties of methyl bromide. Property Synonyms Trade name
Molecular weight, g mole−1 Vapor density, g L−1 Dipole moment, debye Liquid density, g cm−3 20 °C 25 °C Solubility, mg L−1 20 °C 20 °C 25 °C aReid
Value Bromomethane Brom-O-Gas, Meth-O-Gas, Terr-O-Gas 94.94a 3.974e 1.8a 1.676c 1.737a 16,000d 17,500e 13,400b
Property CAS Registry Number Freezing point, °C
Boiling point (at 1.0 atm), °C Koc, cm3/(g %Foc) Log(Kow) Vapor pressure, mmHg 20 °C Henry’s law constant 20 °C 21 °C
Value 74-83-9 −93.7c 3.56e 22b 1.19f 1395 ± 19a,b,d,e 0.23g 0.30 ± 0.02h
et al. (1987): properties handbook; bWauchope et al. (1992); cCRC (1996); dGoring (1962); Index (1996); fSchwartzenbach et al. (1993); gEstimated from mean vapor pressure and solubility data; hGan and Yates (1996). eMerck
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should be noted that because the vapor pressure of MeBr is many times that of chloropicrin, the safety of using low ratios of chloropicrin in the mixture as a warning agent is questionable (van Assche 1971). Methyl bromide is considered to be acutely toxic, with an 8-hr time-weight-averaged limit for human exposure in air of only 5 ppm (ACGIH 1988). Acute toxicity to workers following exposure to its vapor has been a major concern in the many years of MeBr use and is one of the reasons for some early modifications of its application method (e.g., use of surface tarp, mixing with chloropicrin, and use of buffer zones). Fatalities and injuries resulting from exposure to MeBr have been reported, but most incidents are related to structural fumigations rather than soil fumigations (Yang et al. 1995).
III. Methyl Bromide Use as a Soil Fumigant A. History and Scope of Use The insecticidal activity of MeBr was first reported by Le Goupil (1932) and has been subsequently shown on a wide variety of pests including insects, nematodes, rodents, bacteria, viruses, fungi, mites, and weeds (Alexeeff and Kilgore 1983). The early application of MeBr as a fumigant was mainly on stored products (Thompson 1966), but it has been largely replaced by phosphine due to issues of safety and ease of use (Taylor 1994). The concept of fumigating soil with MeBr was introduced later, around the beginning of the 1960s (Hague and Sood 1963), and it has since experienced a steady increase in production and sales. For instance, in 1960, the total MeBr production in the U.S. was 5.7 × 106 kg (Alexeff and Kilgore 1983), which increased to 2.93 × 107 kg by 1992 (UNEP 1995). The total worldwide sales of MeBr increased steadily from 4.56 × 107 kg in 1984 to 7.16 × 107 in 1992 (UNEP 1995). Of the 7.16 × 107 kg of MeBr sold in 1992, 75% was used as a soil fumigant and 15% as a commodity and structure fumigant (UNEP 1995). The U.S., Europe, and Asia are the largest users, consuming 41%, 26%, and 24%, respectively, of the total MeBr in 1992 (UNEP 1995). On a global basis (excluding the U.S.), the predominate use of MeBr as a soil fumigant is in the production of tomatoes (22%), strawberries (14%), and curcurbits (11%) (UNEP 1995). Total methyl bromide preplant use in the U.S. was estimated at ⬃1.6 × 107 kg for the period 1992–1996, most of which was used in the production of tomatoes (30%), strawberries (19%), peppers (14%), grapes (7%), and nursery stock (6%) (Carpenter et al. 2000). Methyl bromide is currently scheduled to be incrementally phased out. In the U.S. and other developed countries that are parties to the Montreal Protocol, MeBr production and importation will be reduced from 1991 baseline levels by 25% in 1999, 50% in 2001, 70% in 2003, and 100% in 2005. Preshipment and quarantine uses are currently exempt from MeBr phase-out. The MeBr phaseout schedule for nonindustrialized countries indicates a 100% reduction by 2015. The MeBr phase-out has produced a wealth of research for chemical and nonchemical alternatives to MeBr fumigation. Research activities are currently
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underway investigating alternative chemicals such as 1,3-dichloropropene (Shaw and Larson 1999; Riegel et al. 2000; Nelson et al. 2001), methyl isothiocyanate (Borek et al. 1997; Shaw and Larson 1999), chloropicrin (Massicotte et al. 1998; Hutchinson et al. 2000), methyl iodide (Eayre et al. 2000; Hutchinson et al. 2000; Zhang et al. 1998), and propargyl bromide (Yates and Gan 1998; Papiernik et al. 2000; Ma et al. 2001), to mention a few. Several nonchemical methods have also been proposed, such as solarization/soil heating (Katan 1992; Porter et al. 1991; Yu¨cel 1995; Stapleton 2000). Brassica roots have been shown to produce isothiocyanates and other allelochemicals, and incorporation of Brassica cover crops may have some efficacy against nematodes and pathogenic fungi (Angus et al. 1994; Gardiner et al. 1999; Walker and Morey 1999). B. Application Methods As a preplant soil fumigant, MeBr is applied by distinctively different methods within the U.S. and outside the U.S. (UNEP 1995). In the U.S., mechanized injection into subsoils is the predominant application method, whereas in many other countries, tarped-surface applications are used. In 1992, about 54% of the overall soil fumigation was conducted with the tarped-subsoil injection method and 40% with the tarped-surface application method (UNEP 1995). The application rate for outdoor mechanical injection is about 24–48 g m−2, whereas a higher rate, normally 45–135 g m−2, is used for the tarped-surface application. A variety of mixtures of MeBr and chloropicrin, with chloropicrin content varying from 0.5% to 37%, are normally used in outdoor fumigation. The formulation used for tarped-surface application in glasshouses is usually a mixture with 2% chloropicrin (UNEP 1995). In outdoor mechanized application, MeBr from pressurized cylinders is injected via tractor-driven shanks or chisels at depths of 20–90 cm into the soil. For shallow injections, the chisels are spaced about 25 cm apart, and the fumigated area is covered with plastic films (almost exclusively low-density or highdensity polyethylene sheets) immediately after injection. The practice of tarping is necessary for controlling weeds near the surface (Lembright 1990), although it also has implications for reducing MeBr emissions, thereby providing better protection for field workers and nearby residents. After liquid MeBr is introduced into the soil, it absorbs heat from the environment around the injection point and rapidly vaporizes. During the initial moments, the distribution pattern of MeBr in the field can be perceived as narrow horizontal cylinders with high concentration of MeBr gas. Driven by the large concentration gradient, MeBr vapor diffuses outward from the lines of injection, and the cylinders of fumigated soil overlap and form uniform concentration profiles in the horizontal plane (Lembright 1990). Plastic covers are removed 2–7 days after fumigation, so planting can begin shortly after the treatment. Untarped deep injection (60–80 cm) is mainly used for eradicating pests for deep-rooted plants such as in orchards and vineyards (UNEP 1995). The spacing between injection lines for deep applications is generally around 1.5–2.0 m.
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In many other countries, manual application of MeBr underneath a presealed plastic tarp is the principal method and essentially the exclusive method for fumigating soil in greenhouses (Maw and Kempton 1973; UNEP 1995). This method is the so-called hot-gas application, in which liquid MeBr is vaporized in a heat exchanger and hot MeBr vapor is delivered through tubing lines with side openings into the space between the plastic tarp and the soil surface (Maw and Kempton 1973; UNEP 1995). Another technique is using liquid MeBr directly instead of the hot gas. Small cans containing MeBr are placed in the surface soil pretarped with plastics, and a special opener is used to release MeBr vapors under the tarp. Following tarped-surface application, MeBr penetrates into deeper soil layers by gas diffusion. The high permeability of MeBr through the traditional polyethylene film was found to be a problem that caused inadequate pest control as well as increased air pollution in the late 1970s. Since then, new materials with reduced permeability have been tested and used for soil fumigation (Kolbezen and Abu-El-Haj 1977; Munnecke et al. 1977; Van Wambeke 1983; de Heer et al. 1983; Daponte 1995; Gamliel et al. 1997). As a result, relatively impermeable plastics have been used on a small scale in some glasshouse fumigations (de Heer et al. 1983; Gamliel et al. 1997).
IV. Ozone Depletion and Methyl Bromide There are both technical and legislative limitations to MeBr use as a soil fumigant. Methyl bromide is toxic to humans, and in most countries its application is restricted to trained applicators, and buffer zones also are required between fumigation sites and nearby residential areas. Fatalities and injuries resulting from MeBr use have been recorded, particularly during fumigation of structures (Yang et al. 1995). Bromide ion residues produced from MeBr in soil have been shown to cause damage to certain sensitive plants such as carnations and wheat (Brown and Jenkinson 1971; Kempton and Maw 1974). In addition, plant accumulation of Br− at excessive levels is considered hazardous for human consumption, and there are maximum limits for Br− concentrations in various fruits and vegetables (Kempton and Maw 1972, 1973; Masui et al. 1978, 1979; Helweg and Rasmussen 1982). These negative effects have influenced some of the application techniques or regulatory restrictions on MeBr use. For instance, leaching fumigated soil to reduce Br− accumulation in plants has been used in glasshouse fumigation in some European countries (Vanachter et al. 1981a,b; Lear et al. 1983; de Heer et al. 1986; Van Wambeke et al. 1988). However, the primary reason that future use of MeBr in agriculture will be eliminated is its involvement in stratospheric ozone depletion. A. Reactions with Ozone It has been known for some time that bromine can catalytically destroy stratospheric ozone (Wofsy et al. 1975; Yung et al. 1980; McElroy et al. 1986; Salawitch et al. 1988; Anderson et al. 1989; Prather and Watson 1990). Reactions
Methyl Bromide
53
involving bromine are believed to be responsible for 20%–25% of the Antarctic ‘ozone hole’ that develops each austral spring (Anderson et al. 1989), which implies that a bromine atom is approximately 40 times more efficient than a chlorine atom in destroying ozone (Wofsy et al. 1975; Salawitch et al. 1988; Solomon et al. 1992). Methyl bromide is unique because it is a significant source of bromine to the stratosphere (Wofsy et al. 1975; Yung et al. 1980; Penkett et al. 1985; Cicerone et al. 1988; Schauffler et al. 1993). However, the case for restricting its use is not clear-cut. Unlike the chlorofluorocarbons (CFCs), atmospheric MeBr is not entirely contributed by human activities. Atmospheric MeBr has abundant natural and anthropogenic sources. Also, its sinks result not only from reactions in the atmosphere but also from interaction with the oceans and land. Thus, estimating the contribution of MeBr fumigation (currently ⬃80% of the entire anthropogenic source) to the depletion of stratospheric ozone is much more complex than it is for other regulated halogenated compounds. To justify the pending suspension of MeBr use in agriculture, it should be established that the known sources of atmospheric MeBr surpass the sinks, and the surplus is contributed by anthropogenic emissions. However, current estimates of global MeBr are out of balance, with sinks exceeding sources by a wide margin (Yvon-Lewis and Butler 1997). The total atmospheric burden of MeBr is believed to be around 145 Gg yr−1 (100–194 Gg yr−1), and the concentration about 10 parts per trillion by volume (pptv), increasing at 0.1–0.3 pptv yr−1 (Khalil et al. 1993; Singh and Kanakidou 1993). The sinks currently thought to remove MeBr from the atmosphere include reactions with OH radicals in the atmosphere (accounting for ⬃86 Gg y−1 MeBr), removal by oceans (⬃77 Gg yr−1), degradation in soil (42 Gg yr−1), and uptake and degradation by plants. The relative strength of each of these sinks is not well quantified. The estimated lifetime of atmospheric MeBr is 0.7 yr (range, 0.4–0.9 yr), with a calculated ozone depletion potential (ODP) of 0.4 (range, 0.2–0.5), according to the World Meteorological Organization 1998 Scientific Assessment of Ozone Depletion (WMO 1999). The known sources of atmospheric MeBr include oceanic emissions, biomass burning, automobile emissions from leaded gasoline, and fumigation. Together, these emissions combine to produce 122 Gg yr−1 of MeBr (range, 43–244 Gg yr−1) (WMO 1999). The 1998 Scientific Assessment of Ozone Depletion estimated oceanic MeBr emissions to be 60 Gg yr−1, with the ocean acting as a net MeBr sink of −21 Gg yr−1 (WMO 1999). Recent research has indicated that the magnitude of the oceanic sink may be −11 to −20 Gg yr−1 (King et al. 2000). Biomass burning (Mano¨ and Andreae 1994) is another significant natural source of atmospheric MeBr, and its contribution is poorly quantified. Global emission of MeBr from biomass burning is estimated to be 20 Gg yr−1 (range, 10–40 Gg yr−1) (WMO 1999). It has also been demonstrated that automobile exhaust from the combustion of leaded gasoline, which contains bromine compounds, can include measurable amounts of MeBr (Harsch and Rasmussen 1977). Emissions from this source could range from 0 to 5 Gg yr−1 (WMO 1999). Additional
54
S.R. Yates, J. Gan, and S.K. Papiernik
potential MeBr sources that have recently been identified include production by plants (Gan et al. 1998a), salt marshes (Rhew et al. 2000), and fungi (Butler 2000). Salt marshes may be a globally important source of MeBr (contributing 7–29 Gg yr−1) (Rhew et al. 2000) and production of MeBr has been observed for a variety of plants (Gan et al. 1998a); therefore, plant sources may account for a large proportion of the “missing source” in current MeBr budgets. Some anthropogenic emissions, such as fumigation of structures, perishables, and durables, are relatively well quantified, because nearly 100% of the applied MeBr is vented into the air during these fumigation processes. The use of MeBr for these fumigations accounts for about 15% of the total production. Trapping and/or decomposing MeBr during structural fumigation can drastically decrease atmospheric emissions of MeBr during these operations. Approximately 85% of the industrially produced MeBr is used as a soil fumigant, equivalent to ⬃65 Gg yr−1 in 1996. The actual discharge of MeBr from fumigated fields into the air is largely determined by the proportion of the applied MeBr that is emitted from the treated soil, which can be reduced through management practices (Wang et al. 1997a; Yates et al. 1998; Gan et al. 1998d). New measurements of the sources and sinks of MeBr are still being actively obtained, as evidenced by many recent reported studies. It is a fact that the relative contribution of MeBr used in fumigation practices is far from well quantified. Despite this, being a significant controllable source, the agricultural use of MeBr becomes a natural target for elimination. It is also assumed that, due to a short atmospheric lifetime of less than 1 yr, the effect of cessation of anthropogenic MeBr emissions on the restoration of stratospheric ozone will be nearly immediate. In comparison, all the chlorofluorocarbons (CFCs) have extremely long lifetimes, and with a complete elimination of emissions of these compounds, it may take many years, or even centuries, to reduce the atmospheric burden of the CFCs to an insignificant level (Butler 1995).
V. Sampling and Analysis of Methyl Bromide in the Air In the course of monitoring MeBr in workplace, field, and ambient atmospheres, sampling and analytical methods of different sensitivities and complexities have been developed. Depending on the sampling device that is used for collecting air samples, MeBr can be either in a contained atmosphere (such as canisters) or adsorbed on a solid adsorbent (such as activated carbon or a porous polymer) before analysis. For the past two decades, quantitation of MeBr has used gas chromatography (GC) exclusively, and electron-capture detectors (ECD) are usually selected over the other types of detectors because of their high sensitivity to halogenated compounds (Scudamore 1988), although very high sensitivity is also found with photoionization detectors (PID) (Dumas and Bond 1985). The main reported sampling and analytical methods of analyzing atmospheric MeBr are summarized in Table 2.
Table 2. Sampling and analytical methods for atmospheric methyl bromide. Sampling method
Tenax-GC Charcoal tube Charcoal tube Charcoal tube Charcoal tube Cold traps Silica capillary Teflon tubing
GC detector
Throughput
Sensitivity
Reference
Cryogenic Cryogenic Cryogenic Direct injection Direct injection
FID ECD FID FID PID
Low Low Low Intermediate Intermediate
Intermediate High Intermediate Low Very high
Jayanty (1989) Gholson et al. (1990) Yagi et al. (1993, 1995) Kolbezen et al. (1974) Dumas and Bond (1985)
Thermodesorptioncryogenic Thermodesorption Solvent extraction Solvent extraction Headspace-GC Headspace-GC
MS
Low
High
Krost et al. (1982)
High Intermediate Intermediate Intermediate Intermediate
Dumas (1982) Eller (1984) Lefevre et al. (1989) Woodrow et al. (1988) Gan et al. (1995b)
Very high Very high
Kallio and Shibamoto (1988) Kerwin et al. (1996)
FID Low FID High ECD (FID) Intermediate ECD Intermediate ECD Very high
Thermo-evaporation ECD Thermo-evaporation ECD
Low Low
Methyl Bromide
Containers Canister Canister Canister Syringe Syringe Adsorbents Polymeric beads
Sample preparation
FID: flame ionization detector; ECD: electron-capture detector; PID: photoionization detector; MS: mass spectrometry.
55
56
S.R. Yates, J. Gan, and S.K. Papiernik
A. Container Methods Steel canisters were used for sampling volatile toxic chemicals in air, such as MeBr, by Jayanty (1989) and Gholson et al. (1990), and good stability and sensitivity were achieved for all the selected analytes. Cryogenic preconcentration was required before the delivery of samples into the GC column. Yagi et al. (1993, 1995) used 500-mL canisters for sampling MeBr to obtain flux measurements under field conditions. Sampling with canisters is labor intensive because the container must be evacuated before sampling, and the contents must be cryogenically concentrated before injection, which limits the number of samples that can be collected and analyzed. Sampling with canisters is therefore not suitable for extensive sampling as needed in volatilization flux measurement under field conditions, although the sensitivity could be very high if a proper detector is used. Using canisters is also not compatible with active (flowthrough) chambers that are used for continuous sampling of the atmosphere. Another container method involves collecting an air sample using a gastight syringe and injecting the contents directly into a GC. In a study of the transport of MeBr in soil after fumigation, Kolbezen et al. (1974) used glass syringes to take and temporarily store soil air samples. The plungers were coated with a film of Triton X-100 to eliminate rapid leakage, and the needle was embedded in a MeBr-impervious sponge. Loss of MeBr was determined to be insignificant within 6 hr, but 5%–7% was lost after 22 hr. The analysis was made by direct injection of the air sample in the syringe into a GC. This method has also been employed in small-scale laboratory experiments (Gan et al. 1998a). B. Adsorbent Methods The most commonly used method for sampling atmospheric MeBr is pumping a relatively large volume of air through one or a series of adsorbent tubes. Methyl bromide in the air stream is trapped in the sample tube containing the solid adsorbent due to its high affinity to the adsorbent. Two types of adsorbent material have been recorded for use with MeBr: activated carbon (charcoal) (Eller 1984; Woodrow et al. 1988; Lefevre et al. 1989; Gan et al. 1995a,b; Majewski et al. 1995; Yates et al. 1996a–c, 1997; Wang et al. 1997a) and porous polymeric adsorbent such as Tenax GC (Brown and Purnell 1979; Dumas 1982; Dumas and Bond 1985; Krost et al. 1982). Activated carbon or charcoal tubes are low in cost (about $1 each), can accommodate large sample volumes, and need minimum preparation before sampling. A typical charcoal tube consists of two adsorption beds: a primary bed (A) and a backup bed (B) in a sealed glass tube. The charcoal can be derived from either coconut or petroleum. Polyurethane spacers are used to separate the two adsorption beds, and a plug of glass wool is usually placed in front of the primary bed to hold the charcoal in the sample tube. Before use, a tube is broken at both ends and then connected to a vacuum source to draw the air to be sampled into the tube. Depending on the sampled volume, air flow rate, and
Methyl Bromide
57
MeBr concentration, multiple tubes connected in series may be required to eliminate loss through breakthrough (Gan et al. 1995a). The number of tubes should be increased when a high flow rate or a long sampling interval is used. Gan et al. (1995a) found, for a single 600-mg coconut charcoal tube at a flow rate of 100 ml min−1, that a sampling interval of ≤2 hr resulted in no breakthrough loss for spiked MeBr masses up to 3.9 mg. Methyl bromide adsorbed in charcoal tubes may be analyzed by two different methods: solvent extraction followed by quantitation in the solvent phase and the so-called headspace-GC method. In solvent extraction, charcoal is transferred into a vial, a known amount of extracting solvent such as carbon disulfide (CS2) is added into the vial, and the vial is sealed (Eller 1984; Lefevre et al. 1989). After the solvent–charcoal mixture is mechanically shaken, an aliquot of the solvent is injected into a GC. This method has the drawbacks of manual sample preparation and presence of other compounds in the final sample solution that may elute with or interfere with MeBr during chromatography (Gan et al. 1995b). This method allows for multiple injections of each sample so that multiple analytes may be measured using different methods or detectors. An alternative method is the headspace-GC method. In headspace-GC analysis, the charcoal is equilibrated with an organic solvent in a closed headspace vial at an elevated temperature for a given period of time, and an aliquot of the headspace containing the analyte is then introduced into the GC column for detection. Benzyl alcohol is often used as the solvent due to its high boiling point (210 °C) (Woodrow et al. 1988; Gan et al. 1995b). When the vial size, solvent volume, and equilibrating temperature and time are fixed, automated headspace injectors give high reproducibility and sample throughput. Gan et al. (1995b) found that the equilibration temperature and time in the headspace autosampler, the size of headspace vials, and the amount of solvent all had an effect on the signal output for a given sample. The sensitivity of analysis can thus be maximized by choosing an optimal combination of these parameters. For instance, to analyze a sample tube containing 600 mg coconut charcoal, the best conditions were determined to be 9-mL headspace vials, 1.0 mL benzyl alcohol, 110 °C equilibration temperature, and 15 min equilibration time (Gan et al. 1995b). Using this method, analysis of a MeBr-containing sample tube takes only 3–4 min, and as many as 300 samples can be analyzed with 24 hr. This method is appropriate when a large number of samples is required, such as when analyzing samples from large-scale field studies measuring MeBr volatilization (Yates et al. 1996b,c, 1997; Wang et al. 1997a). This method has the disadvantage of being destructive, where each charcoal sample can be analyzed with only a single injection. Many kinds of porous polymer adsorbent materials have been used for collecting volatile compounds in the air, including the Chromosorb series, the Porapak series, Ambersorb XE-340, and others. The most popular adsorbent, however, is Tenax-GC, which is a polymer of 2,6-diphenyl-p-phenylene oxide. Brown and Purnell (1979) estimated the safe sampling volume for MeBr to be
58
S.R. Yates, J. Gan, and S.K. Papiernik
0.14 L for sample tubes packed with 0.13 g Tenax-GC. When coupled with a cryofocusing technique, the whole sample can be introduced into the GC column following thermal desorption, which greatly enhances the sensitivity. Detection limits of 500 pg L−1 (Krost et al. 1982) and 35 ng (Dumas and Bond 1985) were reported when this method was used. Compared with charcoal tubes, polymer samplers need to be conditioned before sampling, the safe sampling volume is smaller, the cost is higher, and each analysis takes a longer time. C. Other Methods Other than the container method and the adsorbent method, cryogenic concentration in a cold trap has also been used for collecting MeBr (Kallio and Shibamoto 1988; Kerwin et al. 1996). The cold traps include mixtures of dry ice–acetone, liquid nitrogen, and dry ice–2-propanol. The weakness of this technique is the long time and many steps involved in handling one sample, but it is useful when sample throughput is not a factor and very low detection limits are sought. When extremely high sensitivity is pursued, such as in the case of monitoring MeBr in ambient air, a technique called O2-doping could be useful (Grimsrud and Miller 1978; Kerwin et al. 1996). Grimsrud and Miller (1978) first reported that addition of a fraction of O2 in the carrier gas drastically increased the sensitivity of ECD detection of halogenated methanes including MeBr. When 3%–5% of O2 was added to the carrier gas, signal response for MeBr was enhanced about two orders of magnitude. Using cryogenic concentration and O2-doping, Kerwin et al. (1996) reported a detection limit as low as 0.23 pmol or 22 pg.
VI. Processes Affecting Environmental Fate of Methyl Bromide The efficacy of MeBr fumigation and the extent of volatilization from the soil surface depend on the concentration of MeBr in the soil and the time for which those concentrations are maintained. After MeBr is applied to soil, it is subject to numerous interactions with the soil that may affect its concentration. These processes include transformation and phase partitioning. Transformation also results in the production of Br−, which is considered harmful under certain circumstances. Factors relating to soil and meteorological conditions and the method of application affect MeBr volatilization losses by acting on the transformation and phase-partitioning processes. An analysis of these processes and factors is critical for understanding the fate of MeBr as a soil fumigant, as well as for developing strategies to minimize its volatilization losses. A. Transformation Transformation or degradation of MeBr is an irreversible process that depletes MeBr from the soil–water–air system before it reaches the soil surface and volatilizes into the air. Extremely rapid transformation may deplete MeBr concentrations so quickly that efficacy is compromised. The actual transformation
Methyl Bromide
59
of MeBr in an agricultural soil is the sum of its hydrolysis in water, reactions with soil constituents, and decomposition by soil microorganisms. 1. Water Degradation of MeBr in water is important because it contributes to MeBr degradation in moist soil as well as to its fate in the overall environment. Based on its chemical structure, MeBr is an electrophile, and −Br is reactive as a leaving group and may participate in various nucleophilic substitution reactions (SN1 and SN2 types) in the environment. Water is a weak nucleophile, and therefore hydrolysis of MeBr in water is anticipated: CH3Br + H2O → CH3OH + Br− + H + −
CH3Br + OH → CH3OH + Br
−
Reaction I Reaction II
(1) (2)
The reaction rate constants for reactions I and II are approximately 5 × 10−9 and 10−4 M−1s−1, respectively (Schwarzenbach et al. 1993). In pure water where the OH− concentration is extremely low, reaction I dominates, and the calculated pseudo-first-order half-dissipation time (t1/2) of MeBr should be around 30 d. Mabey and Mill (1978) and Papiernik et al. (2000) report a t1/2 of 20 d, Arvieu (1983) reported a t1/2 of 46 d for MeBr in water at 20 °C, and Gentile et al. (1989) reported t1/2 of 36–50 d in well waters at 18 °C. The relatively slow hydrolysis of MeBr in water was also noted by Herzel and Schmidt (1984). In an attempt to correlate MeBr hydrolysis and pH, Gentile et al. (1992) measured MeBr degradation in buffer solutions with pH 3.0–8.0 and found MeBr hydrolysis rates generally increased with increasing pH. However, in their experiments, they used buffer solutions composed of phosphate and citrate, and apparently nucleophiles other than OH− caused the enhanced hydrolysis in solutions with elevated pH. From the rate constant of reaction II, the MeBr hydrolysis rate should not increase significantly when pH is changed from 7 to 10. In waters that are rich in nucleophiles, such as the supernatant of a salt marsh containing sulfide, MeBr degradation may be accelerated (Oremland et al. 1994a). The reaction produces methanethiol: CH3Br + HS − → CH3SH + Br−
(3)
and further reaction with MeBr produces dimethylsulfide CH3SH + CH3Br → (CH3)2S + Br− + H +
(4)
MeBr was observed to degrade rapidly in anaerobic salt marsh slurries containing sulfide, with a reported transformation half-life ⲏ1 d. Production of methanethiol in slurries doped with sulfide exhibited very rapid reaction, with a MeBr half-life ⬃1 hr (Oremland et al. 1994a). Accelerated transformation by MeBr in aqueous solution containing other nucleophiles (for example, aniline) has also been reported. Reaction with aniline in aqueous solution with a molar ratio of aniline : MeBr of 10 : 1 formed N-methylaniline and N,N-dimethylaniline with a MeBr transformation half-life of 2.9 d (Gan and Yates 1996). The mechanism of photoinduced hydrolysis of MeBr in water was first re-
60
S.R. Yates, J. Gan, and S.K. Papiernik
ported by Castro and Belser (1981). When a pen-ray UV lamp emitting UV at 254 nm was used to irradiate MeBr–water solution in a 4-L closed flask, MeBr was gradually converted to methanol and Br−. The following mechanism was proposed by these authors: CH3Br + hν → (CH3Br)* + H2O → CH3OH + H + + Br−
(5)
Photohydrolysis caused faster dissipation of MeBr in sunlight under sealed conditions (Castro and Belser 1981) and under UV irradiation (Gentile et al. 1989). The significance of reactions with nucleophiles and UV in MeBr transformation in soil water, however, has not been investigated. Hydrolysis of MeBr in aqueous solutions may bear limited significance in determining its fate as a water contaminant. The loss of MeBr in stirred and ventilated waters, or water that had a high surface-to-volume ratio, was found to be very rapid due to volatilization (Gentile et al. 1992). In a study following MeBr kinetics in surface water, Wegman et al. (1981) found that the average half-life for MeBr in surface water at a water temperature of 11 °C was only 6.6 hr. Over its many years of use, contamination of water sources with the parent MeBr has never been a topic of concern. 2. Soil Hydrolysis in water is not the only pathway, and in many cases not even an important pathway, that causes MeBr degradation in soil; this is evidenced by shorter t1/2 values obtained in soil degradation studies, particularly with organic matter-rich soils, as opposed to the t1/2 in water. Two other pathways, i.e., reaction with soil organic matter and microbial degradation, have been identified as contributing to MeBr degradation in soil. Soil organic matter (OM) contains nucleophilic sites such as −NH2, −NH, −OH, and −SH functional groups. Methyl bromide may react with these groups via nucleophilic substitution, as in the hydrolysis reactions: CH3Br + OM − NH2 → OM − NH − CH3 + Br− + H + −
CH3Br + OM − SH → OM − S − CH3 + Br + H
+
(6) (7)
As a result of these reactions, soil organic matter is methylated, and inorganic bromide ion is released. The reaction of MeBr with soil organic matter is supported by the general observation of the close dependence of MeBr degradation on soil organic matter content: in organic matter-rich soils, degradation is consistently more rapid than in organic matter-poor soils (Brown and Jenkinson 1971; Brown and Rolston 1980; Arvieu 1983; Arvieu and Cuany 1985; Gan et al. 1994). For a soil containing 2.81% organic matter, 63 ppm of Br− was generated after exposure to 500 ppm MeBr in closed flasks for 24 hr whereas only 25 ppm of Br− was produced in a soil with 0.93% organic matter (Brown and Jenkinson 1971). Arvieu (1983) studied the rate of MeBr degradation rate in eight soils and found that t1/2 decreased with increasing organic matter content. In a soil with 0.23% organic matter, the t1/2 was 49 d, but in a soil with 5.11% organic matter, the t1/2 was shortened to only 3.6 d. After measuring MeBr degra-
Methyl Bromide
61
dation rates in selected soils, Gan et al. (1994) found that the MeBr degradation rate (based on Br− production) and soil organic matter content were highly correlated, and the measured correlation coefficient (r2) ranged from 0.95 to 1.00. Papiernik et al. (2000) also reported an increase in MeBr degradation rate with increasing soil organic matter content. Spiking soil samples with 14C-labeled MeBr resulted in the formation of nonextractable (bound) residues of MeBr, which increased as extractable MeBr decreased. These bound residues represented transformed (not sorbed) MeBr, as evidenced by the release of equimolar amounts of Br− for each mole of MeBr lost (Papiernik et al. 2000). The reliance of MeBr degradation on soil organic matter content also has implications for MeBr degradation in subsurface layers. Because soil organic matter content normally decreases with increasing depth, MeBr degradation may be much slower in the deep layers, and the overall persistence could be much longer than suggested by the degradation data generated from surface soils. This idea was verified in laboratory incubation studies using soils collected from 0 to 300 cm (Gan et al. 1994). In a Greenfield sandy loam, the t1/2 for MeBr degradation was about 8 d for the 0–30 cm layer, but increased gradually with depth to a t1/2 of 21 d for the 270–300 cm layer. This decrease closely corresponded to the decrease in soil organic matter content (Gan et al. 1994). Biodegradation of MeBr has been documented for isolated bacteria, including the nitrifying bacteria Nitrosomonas europaea, Nitrosolobus multiformis, and Nitrosococcus oceanus (Rasche et al. 1990; Hyman and Wood 1984) and for Methylomonas methanica and Methylococcus capsulatus (Colby et al. 1975, 1977; Meyers 1980; Oremland et al. 1994b). The nitrifier-catalyzed degradation suggests the involvement of ammonia monooxygenase, whereas the consumption of MeBr by the methane-oxidizing bacteria indicates that methane monooxygenases are responsible. Increasing the activity of nitrifying bacteria may increase the rate of biodegradation of MeBr in soil (Ou et al. 1997). Other bacteria capable of degrading MeBr have been isolated from soil. Miller et al. (1997) isolated a facultative methylotroph that could use MeBr as a source of carbon and energy. Oremland et al. (1994b) demonstrated that, at high concentration, biodegradation of MeBr in methanotrophic soils was inhibited due to the toxicity of MeBr itself but became significant at concentrations lower than 1000 ppm. Shorter et al. (1995) suggested that microbial degradation of MeBr at low concentrations (ppb) in surface soils may be important in removing MeBr from the atmosphere, thus reducing its lifetime in the atmosphere and lowering its ozone depletion potential. They observed that MeBr removal from the headspace in closed systems containing soil was more rapid in live soils than in autoclaved soils, and the degradation rate decreased with the depth from which the soil was sampled, which corresponded to the methane oxidation activity of the soil. Hines et al. (1998) reported that at low atmospheric mixing ratios (5 pptv to 1 ppmv) rapid degradation of MeBr was effected by aerobic soil bacteria, resulting in half-lives on the order of minutes for a variety of soils. As the initial concentrations around the injection point are normally on the order of 104 ppmv (Kolbezen et al. 1974) and the MeBr-degrading bacteria are low in population in
62
S.R. Yates, J. Gan, and S.K. Papiernik
normal agricultural soils, bacteria-mediated degradation may be insignificant under typical circumstances. However, studies have indicated that MeBr oxidation can occur in field-fumigated soil. High rates of 14C-MeBr oxidation to 14CO2 were observed in the first few days following soil fumigation where the MeBr concentration was >9.5 µg/g soil (Miller et al. 1997). This oxidation was inhibited by the addition of chloropicrin at concentrations >1.6 µg/g soil. 3. Enhancing Transformation Rates Fumigant degradation in soil varies with soil type and organic matter content (Leistra et al. 1991; Shorter et al. 1995; Gan et al. 1994). However, with reported half-lives of the order of days to months, indigenous soil degradation alone may not be sufficient to significantly affect the emission rate unless amendments are added to soil to enhance degradation. Various organic amendments have been used to increase the degradation rate of MeBr and other fumigants in soil (Gan et al. 1998b,c; Dungan et al. 2002). Gan et al. (1998b) used a biosolid–manure mix and a composted manure to increase the rate of MeBr degradation in soil. They found that the composted manure was more effective in increasing MeBr degradation compared to the biosolid–manure mix and that the degradation rate increased linearly with increasing concentrations of composted manure in the soil. MeBr degradation was nearly 2 times faster than in unamended soil when composted manure was added to soil at a ratio of 1 : 40 dry weight and was 10 times faster at a ratio of 1 : 8 (dry weight). Because similar degradation rates were observed in both sterilized and nonsterilized samples, the increase was attributed to enhanced abiotic reactions. The reactivity of MeBr toward nucleophiles was further demonstrated by its enhanced transformation by thiosulfate salts in solutions and soils. In the reaction, bromide is replaced by thiosulfate (S2O32−), forming a methyl–thiosulfate complex and bromide ion (Gan et al. 1998d). In ammonium thiosulfate (ATS)amended soil, the rate of MeBr degradation was dependent on the level of ATS. At room temperature, MeBr half-life was reduced from 5 d to less than 5 hr when ATS (2 µmole/g) was added to the soil containing MeBr (0.5 µmole/g) (Gan et al. 1998d). It was further shown that after the reaction, the acute toxicity of the aqueous solution substantially decreased (Wang et al. 2000). Thiosulfateinduced transformation occurred at similar rates in different soils, and increased with increasing temperature (Gan et al. 1998d). This reaction has also been used in the development of a simple, safe, and potentially cost-effective method to remediate activated carbon used as a chemical trap in commodity fumigation (Gan and Yates 1998; Yates and Gan 1999). The method will destroy methyl bromide that has been adsorbed to activated carbon. The method only requires that fumigation gases be pumped through a charcoal bed containing a thiosulfate solution, which could be easily accomplished at the fumigation site. 4. Br− Residues in Soil and Plants following Methyl Bromide Fumigation Degradation of MeBr in the soil-water environment produces Br− as the degradation
Methyl Bromide
63
end product, regardless of the transformation pathway. Numerous studies (Table 3) have reported the appearance of elevated concentrations of Br− in soils and plants grown in the soil after MeBr fumigation (Hoffmann and Malkomes 1974). Some of these studies are summarized here to provide a brief review of the problem. Bromide residues produced by MeBr fumigation have importance because excessive uptake of plant materials containing Br− is considered harmful to humans. In addition, some plants, mainly carnations, are sensitive to high Br− levels in the soil. FAO/WHO recommended a maximum acceptable daily intake (ADI) of 1 mg bromine/kg body weight. Different countries and organizations have different tolerance limits in different foodstuffs, but the general limit is 50 ppm for cereals, 20–30 ppm for fresh fruits, 35–300 ppm for dried fruits, and 5–50 ppm for leafy vegetables. The concern for Br− toxicity from edible plant products grown in fumigated fields was the main reason for the suspension of MeBr in Germany (Anonymous 1980), the adoption of more restrictive regulations such as lower doses and leaching with water after fumigation in the Netherlands in the early 1980s, and the ultimate ban in the Netherlands in the early 1990s (UNEP 1995). The production of Br− from MeBr in fumigated soil is a result of the transformation of MeBr in soil. All known and proposed mechanisms of MeBr transformation in soil and water release Br−. All MeBr that is not volatilized from the soil surface will be converted to Br− in the soil. Thus, concentrations of Br− resulting from MeBr fumigation are dependent on the factors that impact MeBr degradation and volatilization and on the rate and extent of Br− removal following its formation. These factors include methods of application, application rates, surface cover removal time, soil organic matter content and texture, permeability of surface tarp, irrigation or precipitation after fumigation, and timing of successive fumigations. Masui et al. (1978, 1979) fumigated soil beds with incremental dosages of MeBr under vinyl film-tarped conditions, and found that soil Br− levels generally increased with increasing dosages in the same soil but that 10 times more Br− was formed in a soil that had a higher clay content. Nazer et al. (1982) found that soil Br− concentration increased proportionally with application rates from 45 to 135 g m−1. Hass and Klein (1976) reported that more Br− was formed in the soil when it was fumigated with liquid MeBr than with gaseous MeBr, and organic amendment greatly stimulated the conversion of MeBr to Br−. Fransi et al. (1987) noted, after fumigation at 90 g m−2 in the field, that Br− residues ranging from 5 to 10 ppm were distributed to a depth of 50–60 cm, where a compacted layer existed. In packed soil columns treated with MeBr under polyethylene film-tarped conditions, Vanachter et al. (1981a) found that nearly three times more Br− was produced in a loamy soil with 7.22% organic matter than in a sandy soil with 2.15% organic matter. In outdoor broadcast fumigation (Yates et al. 1996a; Yagi et al. 1995), because typically a much lower rate is used, Br− production in the soil profile is significantly lower than that found in fumigated glasshouse soils. After a tarped, shallow (25–30 cm) injection at 24 g m−2, noticeable increases of Br− were
64
Table 3. Br− residues in soil and plant tissues after soil fumigation (in mg kg−1). Fumigation typea
Dosage (g m−2)
Soil typeb
Soil Br−c (mg kg−1)
Plant tissued
96
Unknown
⬃12
Wheat leaves
Tarped-PE
24 49 98 24
Unknown
Lettuce
24
Loam/peat (7 : 3)
Tarped-PE (gas)
24 73 24 73 24 73 73
Loam Loam Loam/peat (4 : 1) Loam/peat (4 : 1) Loam/peat (1 : 4) Loam/peat (1 : 4) Sandy loam
⬃2 ⬃3 ⬃5 ⬃17 ⬃14 ⬃12 ⬃18 ⬃18 ⬃16 ⬃10 ⬃23 ⬃13 ⬃32 ⬃20 ⬃52 ⬃33
Tarped-PE
73
Sandy loam
⬃35
Tarped-PE (2 d) (+ discing) (+flooding) Tarped-PE (5 d) (+ discing) (+flooding) Tarped-PE
Loam/peat (7 : 3)
Carnation
Carnation
Tomato Tomato Tomato Tomato
leaves fruit leaves fruit
Reference
3500–6100 (1st yr) 2400–4200 (2nd yr) 460–2500 (3rd yr) ⬃430 ⬃770 ⬃870 1100 1300 800 1800 1800 1600
Brown and Jenkinson (1971)
9300 720
Kempton and Maw (1972)
Kempton and Maw (1973)
Kempton and Maw (1974)
S.R. Yates, J. Gan, and S.K. Papiernik
Tarped-PE
Plant Br− (mg kg−1)
Table 3. (Continued).
Fumigation
typea
Tarped-vinyl
Tarped-vinyl
Tarped-PE Tarped-PE (gas)
PE: polyethylene.
Loam
Clay
Clay (Iwata)
Clay (Takamatsu)
Unknown
Soil Br−c (mg kg−1) ⬃9 ⬃7 ⬃9 ⬃4 ⬃6 ⬃12 ⬃6 ⬃12 ⬃4 ⬃12 ⬃4 ⬃12 ⬃19 ⬃73 ⬃165 14–27
90
14–49
135
14–62
75 50 75 50 80
Clay
Sandy
⬃17 ⬃12 ⬃12 ⬃5 ⬃32
Plant tissued Muskmelon fruit Muskmelon fruit Muskmelon fruit Cucumber fruit Cucumber fruit Cucumber fruit Tomato fruit Pepper fruit Eggplant fruit Strawberry fruit
Tomato fruit (wet) Cucumber fruit (wet) Tomato fruit (wet) Cucumber fruit (wet) Tomato fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Cucumber fruit (wet) Tomato fruit
Plant Br− (mg kg−1) 164 303 987 97 98 189 118–136 162–196 98–312 214–383 197–229 360–368 164–195 167–255 228–328 6–11 8–11 8–13 9–14 8–13 13–17 ⬃25 ⬃15 ⬃19 ⬃17 55 (wet)
Reference Masui et al. (1978)
Masui et al. (1979)
Nazer et al. (1982)
Malathrakis and Sarris (1983)
Fallico and Ferrante (1991)
65
Tarped-PE
200 400 600 200 400 600 100 200 100 200 100 200 100 200 300 45
Soil
typeb
Methyl Bromide
Tarped-PE
Dosage (g m−2)
66
S.R. Yates, J. Gan, and S.K. Papiernik
found from the surface to about 100 cm below surface (Yates et al. 1996a). The highest Br− concentrations were found near the soil surface, and about 39% of the injected MeBr was transformed to Br−. Drastic variations were noticed in the distribution of Br− in soil, and numerous soil cores had to be sampled to obtain a meaningful average (Yates et al. 1996a). These variations could be caused by the nonuniform distribution of MeBr in soil or the effect of soil heterogeneity in degrading MeBr. The accumulation of Br− in plant tissues is usually proportional to the Br− concentration in the soil. Therefore, factors affecting the production and retention of Br− in the soil around plant root systems also affect Br− uptake by the plant (Lavergne and Arvieu 1983). In addition, the accumulation and distribution of Br− in a plant is also affected by plant species, age, and position of the organs. Brown et al. (1979) established a linear correlation (r2 = 0.74) between Br− concentrations in soil and barley plants from numerous plots. They also found that the Br− level in plants varied from one location to another and from one species to another. Soil texture, temperature, leaching, and nonuniform distribution of MeBr were among the other factors that might have caused the observed variations. Kempton and Maw (1972) reported that the accumulation of Br− by lettuce increased with application rates from 49 to 98 g m−1, and Br− residues as high as 10,100 ppm were recovered in the dried tissue. Staerk and Suess (1974) found that in Germany, the natural bromine content in fresh vegetables varied from 0.5 to 3.0 ppm. After the treatment of soil with MeBr, the bromine content of the vegetables increased by factors of 200 or more. They found that the bromine uptake was influenced by several parameters such as plant species, soil type, MeBr dose, vegetation time, and rotation. In England and Wales, lettuce grown on unfumigated soil contained less than 10 ppm Br−, while most lettuce grown on MeBr-fumigated soil contained higher levels of Br−, with a proportion in excess of 1,000 ppm (Roughan and Roughan 1984a,b). Basile and Lamberti (1981) reported that tomato, string beans, radish, egg plant, zucchini, and pepper grown in MeBr-treated soil (60 g m−2) all had markedly higher Br− concentrations than those grown in untreated soil. The accumulation of Br− also appeared to be related to the interval between soil fumigation and planting as well as to the frequency of fumigation (Roughan and Roughan 1984a,b). White and Hunt (1983) found that fumigating seedbeds with MeBr gave high Br− concentrations in the seedlings of cabbage, cauliflower, and Brussels sprouts, but 6 wk after transplanting, Br− level in the plants decreased to nearbackground level. Kempton and Maw (1973) found that uptake of Br− by tomato plants was directly influenced by fumigation rates and the interval between fumigation and planting. The concentration in the leaves decreased from the base to the tip of the plants and increased with the age of the tissue, but the accumulation in fruits was less than that in the leaves (Kempton and Maw 1973). Very high Br− concentrations derived from MeBr fumigation also exhibited phytotoxicity to certain plants. Brown and Jenkinson (1971) noticed that wheat plants were discolored (“scorched”) as a result of MeBr fumigation, particularly near the injection sites. They associated this phenomenon with Br− uptake by
Methyl Bromide
67
the plants, and found as high as 6100 ppm Br− in the scorched plants. Kempton and Maw (1974) found that carnations were extremely sensitive to inorganic Br−, and many carnation plants were injured or even killed when planted in MeBr-fumigated soil. Flooding the soil or incorporating peat into soil alleviated the injury due to their effects on lowering soil Br− level. Possible Br− residues in edible plant parts is an important objection against the use of MeBr as a soil fumigant in some European countries, where MeBr is often repeatedly used in glasshouses. Various methods have been studied to reduce Br− uptake by plants. Among the tested methods, the most effective approaches are leaching the fumigated soil with abundant amounts of water and reducing the application dosage by using less-permeable tarps or mixtures of MeBr with a secondary fumigant. In packed soil columns, the amount of Br− leached out of soil was found to be directly dependent on the amount of water added, and the amount of water needed was dependent on the soil type (Vanachter et al. 1981a,b). After leaching with 40 cm water, 47% of the produced Br− was removed from a sandy soil column and 92% was eliminated from a loamy soil column. MaCartney and Price (1988) reported that after flooding the treated soil with 15 cm of water, the Br− content in the surface soil layers decreased from 34 to 4 ppm. Roorda et al. (1984) also found that leaching effectively reduced soil Br− derived from fumigation, and in some cases, soil Br− level became lower than that before the fumigation. In addition to leaching, application methods that allow a lower input of MeBr into the soil were found to result in lower Br− levels in the fumigated soil. Coosemans and Van Assche (1976) compared soil Br− production following applications of pure MeBr and a mixture of 70% MeBr and 30% chloropicrin and found that soil Br− concentrations were considerably lower with the application of the mixture. There are some negative effects associated with leaching. Trace amounts of MeBr were carried in greenhouse leachate water into surface water when leaching was started directly after the films were removed (1 wk after the fumigation). As a result, some small fish were killed near the drainage discharge point (Wegman et al. 1981). The concentration of MeBr in drainage water was found to be related to the covering time and the time that irrigation started (de Heer et al. 1986). Leaching fumigated greenhouse soil also resulted in high Br− concentrations in surface and drainage waters near the fumigation site, but it was found that the Br− was rapidly diluted after leaching was stopped (Wegman et al. 1981; Vanachter et al. 1981b; de Heer et al. 1986). Calculation of Br− mass balance over a year in a polder district of the Netherlands revealed that fumigation with MeBr contributed 68% of the total Br− found in the precipitation, surface water, and groundwater, which corresponds to conversion of 14% of the applied MeBr (Wegman et al. 1983). 5. Production of MeBr by Plants Production of MeBr by plants has been demonstrated in laboratory studies (Saini et al. 1995; Gan et al. 1998a). Floating plant leaf disks on solutions containing high concentrations of halide ions resulted in evolution of MeBr and other methyl halides (Saini et al. 1995). The
68
S.R. Yates, J. Gan, and S.K. Papiernik
capability of plant leaves to methylate halide ions and release methyl halides appears to be common among higher plants. A high proportion of species tested demonstrated the capability to methylate halide ions, with the family Brassicaceae being particularly efficient (Saini et al. 1995). Plants may contain a variety of methyl transferase enzymes that effect this reaction. Intact plants grown in soil containing Br− were also observed to produce MeBr. In a study by Gan et al. (1998a), several species from the Brassicaceae family were grown in soil containing 0.4–100 mg/kg Br−. Production of MeBr per gram of plant mass was correlated to the soil Br− concentration. No MeBr production was observed in Br−-treated soil without plants. No MeBr was observed in microcosms from which the aboveground plant mass was removed, indicating that MeBr production or release could not be caused by plant roots alone. Because low concentrations of Br− are ubiquitous in soils, terrestrial plants may be an important component of the MeBr cycle. Flux of MeBr and other methyl halides from rice paddies was measured by Redeker et al. (2000). Results indicated production of MeBr by rice paddies, although the study did not include a plant-free control, so it was not verified that plants were responsible for MeBr production. The flux of MeBr depended on the rice growth stage, and maximum MeBr flux occurred during heading and flowering. Plots in which the rice straw was incorporated into the soil had increased soil organic matter content and increased MeBr production. At one site, MeBr production was proportional to soil Br− concentrations. Muramatsu and Yoshida (1995) also observed production of methyl halide (methyl iodide) by rice plants, with seasonal variation in MeI production. Unplanted soil controls also produced measurable MeI, but both flooded and unflooded soils produced less MeI than the treatments containing rice plants. Gan et al. (1998a) estimated a global production level of MeBr by rapeseed and cabbage plants to be 6.6 ± 1.6 Gg/yr and 0.4 ± 0.2 Gg/yr, respectively. Rice paddies may contribute approximately 1.3 Gg/yr (Redeker et al. 2000). These values are of the same order of magnitude as some anthropogenic sources of MeBr (Yates et al. 1998). It seems likely that the contribution of plants is important in the MeBr cycle, although it is often not accounted for in current MeBr budgets. B. Phase Partitioning 1. Henry’s Law Constant It has been long established that the movement of a volatile chemical in soil is controlled by its distribution behavior over the soil–water–air phases. The reported Kh for MeBr at 20 °C varies from 0.24 to 0.30 (Siebering and Leistra 1978; Gan and Yates 1996), and changes with temperature. With Kh of this magnitude, it can be expected that the movement of MeBr in unsaturated soil is mainly driven by its diffusion via the vapor phase (soil air). The temperature dependence of the Henry’s law constant for MeBr is shown in Fig. 1, including Arrhenius equations and fitted parameters. 2. Adsorption The other distribution factor, the adsorption coefficient, Kd (mL g−1), is important as a retaining force in slowing down MeBr transport through
Methyl Bromide
69
Fig. 1. Vapor pressure, solubility, and Henry’s constant as a function of temperature.
70
S.R. Yates, J. Gan, and S.K. Papiernik
the soil. There are a few published measured or estimated Kd and Koc values for MeBr. The reported Koc ranges from 9 to 22 (Briggs 1981; Karickhoff 1981; Rao et al. 1985), which corresponds to a Kd of 0.09–0.22 in a soil with 1% organic carbon. Arvieu (1983) measured MeBr adsorption and desorption and found different characteristics for soil with different organic matter contents. In organic matter-poor soils, the adsorption of MeBr is very weak unless the soil is very dry. In organic matter-rich soils, the adsorption is considerably greater. The same author also noted that the adsorbed MeBr became resistant to desorption. Gan and Yates (1996) observed that degradation of MeBr during the equilibration in adsorption studies might have contributed to the observed increased adsorption in soils with high organic matter content. A noticeable fraction of the spiked MeBr was degraded to Br− during a 16-hr shaking period in organic matter-rich soils. This phenomenon may be also responsible for the irreversibility found in MeBr desorption isotherms (Arvieu 1983). After correcting for the degraded fraction, MeBr adsorption became negligible in all the tested soils (Gan and Yates 1996). Therefore, MeBr can be considered to be a nonadsorbing chemical in soil with normal water content.
VII. Simulating the Environmental Fate of Methyl Bromide Models of various complexity are available to simulate the transport of water, heat, and MeBr in variably saturated soils. A comprehensive simulation requires the use of conservation (i.e., mass or energy balance) equations as well as other equations that describe the mass or energy flux, reactions, and interactions that are important to describe the problem. Under natural conditions existing in agricultural soils, many factors affect the transport process to varying degrees, making it virtually impossible to use models for reliable prediction in all cases. The fate and transport of MeBr is strongly affected by temperature. To conduct an accurate simulation under temperature-dependent conditions, a governing equation is necessary that describes the transport of energy. In general, this equation will be coupled to both the water flow and chemical transport equations and considerably increases the model and solution complexity. A. Transport Model A common approach for simulating the fate and transport of MeBr for saturated and unsaturated water flow conditions, with consideration of variable soil temperature, includes descriptions for at least three governing processes: water flow, heat transport, and fate and movement of MeBr. Programs exist that numerically solve the nonlinear partial differential equations for one- and two-dimensional systems, nonequilibrium coupled transport of water, heat, and solute (in both liquid and gaseous phases) in a variably saturated porous medium. Degradation is usually described using a first-order decay reaction, and often the degradation rate in each phase (liquid, vapor, and solid) can be specified. The governing transport equations can be written as follows (S˘imu˚nek and van Genuchten 1994).
Methyl Bromide
Water Transport:
71
册
冋
∂h ∂θ ∂ = Kij(h) + Kiz(h) − S ∂t ∂xi ∂xj
(8)
where θ is the volumetric water content [L3 L−3], h is the pressure head [L], Kij are components of the unsaturated hydraulic conductivity tensor [L t−1], and S is a sink term [t−1]; t is time, x is distance [L], and indices i and j represent the horizontal and vertical directions. Heat Transport: Ch(θ)
册
冋
∂T ∂ ∂T ∂T = λij(θ) − Cwqi ∂t ∂xi ∂xj ∂xi
(9)
where Ch and Cw are the volumetric heat capacity for the porous media [J m−3 K−1] and liquid, respectively, and λij is the apparent thermal conductivity [Wm−1 K−1]. Solute Transport:
冋
册 冋
册
∂θCL ∂ρbCS ∂ηCg ∂ ∂CL ∂Cg ∂ + + = θDijw + ηDgij ∂t ∂t ∂t ∂xi ∂xj ∂xi ∂xj +
(10)
∂qiCL − (µwθCL + µgηCg + µSρbCS) − SCr ∂xi
where CL [M L−3], CS[M M−1], and Cg[M L−3] are solute concentrations for the liquid, solid, and gaseous phases, respectively; q is the volumetric flux density; µw , µs , and µg are first-order rate constants [t−1] for solutes in the liquid, solid, and gas phases, respectively; θ is the volumetric water content, ρ is the soil bulk density, η is the soil air content, S is the sink term in the water flow equation [t−1], Cr is the concentration of the sink term, Dijw is the dispersion coefficient tensor for the liquid phase [L2 t−1], and Dgij is the diffusion coefficient tensor for the gas phase. Numerous computer programs have been developed to evaluate the effects of interacting processes and factors on pesticide movement through the root zone and to the groundwater. The approach used in developing the programs varies with the intended use of the model. Some of these include GLEAMS (Leonard et al. 1987), LEACHM (Wagenet and Hutson 1987), PRZM (Carsel et al. 1985, 1998), PESTAN (Enfield et al. 1982), and SESOIL (Bonazountas and Wagner 1984). Some of these models are not capable of predicting pesticide movement when water is applied in a controlled manner by furrow or subsurface drip irrigation systems. This limitation has led to the development of processbased models that can be used to predict the transport in irrigated agriculture: CHAIN-2D (Sˇimu˚nek and van Genuchten 1994), HYDRUS-2D (Sˇimu˚nek et al. 1996), and PESTLA (van den Berg and Boesten 1997). 1. Volatilization Boundary Condition For methyl bromide, critical processes affecting the fate and transport in soils are vapor diffusion and volatilization.
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S.R. Yates, J. Gan, and S.K. Papiernik
Volatilization is an especially important route of dissipation because of MeBr’s large vapor pressure and Henry’s law constant, as demonstrated in recent field experiments (Yagi et al. 1995; Majewski et al. 1995; Yates et al. 1996b). Excessive volatilization is associated with many problems, such as a reduction in the amount of material available to control pests and increased potential for contamination of the atmosphere. Emission losses to the atmosphere pose an increased risk to persons living near treated fields. When simulating MeBr emissions to the atmosphere, the approach used to describe the soil surface– atmospheric boundary condition strongly affects the simulated emission response. The most common volatilization boundary condition used in current models is based on stagnant boundary layer theory (Jury et al. 1983). This approach assumes that a thin stagnant air layer occurs at the soil–atmosphere interface and that chemical movement across the layer is the result of vapor diffusion. The controlling parameter is the mass transfer coefficient, which is expressed as the ratio of the binary diffusion coefficient (i.e., air and MeBr) to the boundary layer thickness (Jury et al. 1983). A limitation of this approach is the estimation of the thickness of the stagnant boundary layer. Further, for some atmospheric conditions (e.g., changes in barometric pressure), it is likely that chemical transport occurs by both advection and diffusion. For these situations, assuming a stagnant boundary layer is inappropriate and more complex boundary conditions are required (Massmann and Farrier 1992; Chen et al. 1995). An advantage of the stagnant boundary layer approach is that information about atmospheric conditions is unnecessary. However, adopting this boundary condition produces MeBr emission histories that are very regular and often do not resemble the erratic behavior commonly observed in the field (Majewski et al. 1995; Yates et al. 1996b, 1997). A more accurate description of the volatilization process requires the coupling of soil-based processes with those operating in the atmosphere, which led Baker et al. (1996) to develop an alternate formulation for the mass transfer coefficient that includes atmospheric resistance terms. This boundary condition depends on several aerodynamic parameters, such as the roughness Reynolds number, the Schmidt number, the friction velocity, the wind speed, and an atmospheric stability term. The atmospheric resistance to diffusion near the soil surface and aerodynamic resistance from the diffusive layer to the measurement height affects the predicted emission rate. Further research is needed to evaluate the effectiveness of this approach in simulating the volatilization boundary condition, especially for agricultural fumigation. Several studies have been conducted to determine whether conventional modeling approaches can accurately predict the rate of MeBr volatilization from bare soils (Wang et al. 1997b; Yates et al. 2002). Wang et al. (1997b) used CHAIN-2D to simulate methyl bromide emissions from a 3.5-ha field and compared the simulation results to emissions measured in a field experiment (Yates et al. 1996c). They found that the model simulated the total emission within a few percent of the measured value but the pattern of instantaneous emission rate
Methyl Bromide
73
was much more regular than the measured values and, at times, a value of the predicted volatilization rate could be very different from the measured value. Yates et al. (2002) conducted a similar study using the same experimental data and the volatilization boundary condition of Baker et al. (1996). They found that the predicted emissions had a more realistic temporal pattern compared with a simulation based on the stagnant boundary layer. The total emissions were also within a few percent of the measured value. When discrepancies occur, it cannot be determined whether the model or the measured values were in error because the measured volatilization rate is also subject to uncertainty (Majewski 1997). Further research is needed to improve the accuracy of volatilization measurements and simulation models. Research is also needed to develop and test methods for coupling atmospheric and soil processes in models so that more accurate predictions of the volatilization rate can be obtained. B. Mobility Indices Along with highly sophisticated numerical approaches, there is a need for simple methods for ranking pesticides by their potential to contaminate the environment. Jury et al. (1983, 1984a,b) described a screening model for assessing the relative volatility, mobility, and persistence of pesticides and other trace organic chemicals in the soil. Other mobility indices reported in the literature include the retardation and attenuation factors (Rao et al. 1985) and the convective and diffusive mobility times (Jury et al. 1984b). The retardation factor (RF) is an index of the relative time needed for a pesticide to move past some specified depth compared to a nonadsorbing tracer. The attenuation factor (AF) is the fraction of pesticide mass that is likely to move past some specified depth and includes the effects of adsorption and degradation. The convective and diffusive mobility times (tc and tD) give a measure of the time needed for a pesticide to travel a specified distance by convection or diffusion, respectively. This information is important in determining the suitability of candidate pesticides and in developing management practices. Jury et al. (1984b) classified MeBr as one of the most mobile compounds considered, with a class 5 convective and class 3 diffusive mobility. In terms of persistence, MeBr was found to be class 5, very short lived. Mobility indices were used by Yates and Gan (1998) to compare the transport potential of MeBr to several alternative fumigants, herbicides, and insecticides (Table 4). The results demonstrate that MeBr is highly mobile, an ideal characteristic for a fumigant, but it is somewhat more persistent in soil than the other soil fumigants. In general, all the fumigant materials, with the possible exception of methyl isothiocyanate (MITC) have high diffusive mobility and are quickly attenuated in soil. The mobility indices can provide insight into observed fumigant behavior. For example, some fumigants (e.g., MITC) have relatively low diffusive mobility compared to MeBr and often suffer from difficulties achieving a uniform soil concentration throughout the treatment zone. Olson and No-
74
S.R. Yates, J. Gan, and S.K. Papiernik
Table 4. Mobility indices for MeBr and selected alternative fumigants and herbicides.
Pesticide
RF
AF R = 25 cm
tc R = 10 cm
tD R = 10 cm
Class
Methyl bromide 3-Bromopropyne (Z)-1,3-D (E)-1,3-D Methyl isothiocyanate 2,4-D (acid) Atrazine Carbaryl Lindane
2.37 1.25 2.81 2.79 1.34 2.1 6.52 17.6 61.7
0.59 0.18 0.042 0.019 0.37 0.33 0.56 0.0001 0.45
7.12 3.75 8.44 8.38 4.01 6.31 19.62 52.69 185.19
4.03 11.54 16.19 27.6 55.18 —a —a —a —a
3 3 3 2 2 1 1 1 1
RF: retardation factor; AF: attenuation factor. aGreater than 105.
ling (1994) found that MITC may not completely control root gall in tomatoes and strawberries because of poor chemical distribution, causing others to develop better methods for distributing MITC in soil (Juzwik et al. 1997; McHenry 1994). MeBr has relatively high diffusive mobility and low retention on soils and therefore provides a very uniform concentration distribution and pest control. C. Simulating Transport in Relatively Dry Soils For certain situations, it is possible to simplify the mathematical model given by Eqs. 8 to 10. Within the context of predicting MeBr transport in the subsurface under typical application conditions, it is often reasonable to neglect the flow of water. For shank injection, this can be a reasonable assumption because the fumigation occurs in relatively dry soils. For dry-soil conditions, the magnitude of the effective diffusion coefficient is determined primarily by the vaporphase component (Fig. 2); this is a result of the magnitude of the gas-phase diffusion coefficient, which is approximately four orders of magnitude greater than the liquid-phase diffusion coefficient (Jury et al. 1983). Further, given the short time scale of a soil fumigation, water movement (redistribution) is also very small. The MeBr transport process can be approximated mathematically with a single transport equation that is not coupled to the Richard’s water flow equation. Such an assumption would not be appropriate if the fumigant were applied in moist soils or delivered in irrigation waters. 1. Using Models to Extrapolate Laboratory Data to Field Situations Comparing experimental measurements from laboratory and field studies can be problematic. In the laboratory, soil columns are often of limited length and have impermeable surfaces at the lower end of the column. In fields, there are generally no impermeable boundaries and MeBr gases can diffuse unimpeded. There-
Methyl Bromide
75
Fig. 2. Contribution of vapor and liquid components to the effective soil diffusion coefficient as a function of water content. For water contents below 0.375 (dotted line), the vapor-phase component dominates. Soil porosity = 0.4. Devap and Deliquid use the % of De scale.
fore, deep diffusion in the field promotes additional soil residence time (and concomitant degradation) and reduces emissions. This effect was illustrated by Gan et al. (1997a), who found that total MeBr emission values from an untarped 60-cm laboratory column could overpredict field values by 11%–58%, respectively, for a 20-cm and 60-cm injection depth. For a 60-cm injection, the measured value of 60% considerably overestimated a field measurement of 21% (Yates et al. 1997). This finding led to the development of a method to extrapolate laboratory measurements conducted in soil columns to values more representative of field conditions. The corrected emissions were obtained by using two analytical solutions to a partial differential equation describing vapor diffusion in a one-dimensional homogeneous soil system. For both, it was assumed that the chemical and material properties are spatially and temporally constant. These assumptions allow simulation using a simple differential equation: η
∂Cg ∂Cl ∂Cs ∂2Cg ∂2Cl +θ + ρb = Dg + D − ηµgCg − θµlCl − ρbµsCs (11) l ∂t ∂t ∂t ∂x2 ∂x2
where Cg, Cl, and Cs are the vapor-, liquid-, and solid-phase concentrations [mg/ m3], Dl and Dg are the liquid and gaseous diffusion coefficients [m2/d], and θ, η, ρb, and µi are the water content, air content, bulk density, and degradation coefficient for the ith phase. The three terms on the left-hand side of Eq. 11 describe the time rate of change in the gas-, liquid-, and sorbed-phase concentration. On the right-hand side of Eq. 11, the first two terms describe gas- and liquid-phase diffusion and the remaining terms describe first-order degradation in each phase.
76
S.R. Yates, J. Gan, and S.K. Papiernik
For both analytical solutions, linear relationships are used to characterize the partitioning between the solid and liquid phases and for liquid–vapor partitioning: C s = K d Cl C g = K h Cl
(12)
with initial and boundary conditions ∂C(0, t) Surface Boundary = h [C(0, t) − Cair] ∂x ∂C(x, t) (13) *x→L = 0 Lower Boundary ∂x C(x, 0) = Co [u(x − xo − δ) − u(x − xo + δ)] Initial Condition Ds
where Ds is the effective soil diffusion coefficient [m2/d] and h is a mass transfer coefficient [m/d] that characterizes the resistive nature of the interface between soil and the atmosphere. The initial condition assumes that the chemical is injected into the column at a specified depth, xo, with a pulse width of 2δ. For a laboratory column with an impermeable barrier at the bottom, the solution to Eq. 11, given the initial and boundary conditions in Eq. 13, is this: C(x, t) = e−µt
∞
∑a
2
n
e−knDst Cos[kn (F − x)]
(14)
i=1
where an =
4Co{Sin[kn (F − zo + δ)] − Sin[kn (F − zo − δ)]} 2knF + Sin(2knF)
(15)
where δ is some small interval and kn are the roots of the following equation: kn → h Cos[F kn] − Dskn Sin[F kn] = 0
(16)
There are an infinite number of values for kn that satisfy Eq. 16. A second solution for transport in a homogeneous field soil that does not have an impermeable vapor barrier at depth (i.e., when L → ∞ in Eq. 13) is this:
冋 册 冋 册 冋√ 册 冋√ 册 冋 √ 册 冋 √ 册
z + z2 z − z2 2 C(x, t) µ t e = Erfc + Erfc Co √4Dst √4Dst − Erfc
z + z1
4Dst
− Erfc
+ 2 eh[ht+(z+z1)]/Ds Erfc − 2 eh[ht+(z+z2)]/Ds Erfc
where z1 = F − zo − δ and z2 = F − zo + δ.
z − z1
4Dst
2ht + z + z1 4Dst
2ht + z + z2 4Dst
(17)
Methyl Bromide
77
Equations 14 and 17 can be used to extrapolate laboratory measurements to analogous field values by finding the transport parameters that produce the best fit of Eq. 14 to the laboratory measurements, then using these parameters in Eq. 17 to obtain corresponding corrected (or “field”) values. Values obtained through this sort of procedure should be more representative of field values compared to the original laboratory measurements. An application using this method is discussed in Section X.A.1.
VIII. Methyl Bromide Diffusion in Soils A. Diffusion Coefficient The value of the diffusion coefficient of a chemical in the vapor phase is generally 104 times larger than that in the liquid phase (Jury et al. 1983). The diffusion coefficient can be estimated using a variety of methods (Reid et al. 1987), including the Fuller correlation Dab =
0.00143T 1.75
冋
1/3 1/3 P M1/2 ab (Σν)a + (Σν)b
册
2
(18)
where Dab is the binary diffusion coefficient (cm2/s), T is the absolute temperature (K), Mab = 2/(M a−1 + Mb−1), Ma and Mb are the molecular weights of air and MeBr, respectively, P is the pressure (bars), and Σν is obtained using the atomic diffusion volumes (Reid et al. 1987). Using Eq. 18 yields an estimated diffusion coefficient for MeBr of 0.114 cm2 s−1 at 20 °C and 1 atmosphere ambient pressure. The temperature dependence of the diffusion coefficient, shown in Fig. 3, appears to be nearly linear over the temperature range 0°–60 °C. The temperature dependence of the binary diffusion coefficient can be described using Eq. 18 or using activation energy and the Arrhenius equation as shown in Fig. 3. Using a screening model, Jury et al. (1991) found that the movement of a chemical is dominated by vapor-phase diffusion if the air-to-water partition coefficient, or the Henry’s law coefficient (Kh) is Ⰷ10−4. As the Kh for MeBr is approximately 0.25, transport in the vapor phase is important in describing the fate and transport in soil. B. Methyl Bromide Diffusion in Soils Because of its large air–water distribution coefficient (e.g., Kh) and insignificant adsorption to soil, a large fraction of MeBr partitions into the soil air in unsaturated soils. For instance, in a soil with 20% volumetric water content and 20% air space, as much as 24% of the total MeBr may be present in the soil air with negligible adsorption onto the soil. This partitioning forms the basis of rapid vertical movement of MeBr immediately following a typical subsoil injection, and the potential for significant volatilization losses into the air if movement
78
S.R. Yates, J. Gan, and S.K. Papiernik
Fig. 3. The temperature dependence of the binary diffusion coefficient for MeBr in air. The solid line is from Eq. 18, and an equivalent Arrhenius equation is shown.
across the soil surface is not suppressed. Figure 4 contains a series of concentration curves measured at different time intervals following an injection at 30 cm into a homogeneously packed Greenfield sandy loam column (Gan et al. 1997a). It is clear that after the injection MeBr quickly expands and moves vertically in both directions. Unless the soil surface is sealed with a relatively impermeable
Fig. 4. Methyl bromide distribution in the soil gas phase after injection at 30 cm. Soil surface tarped with 1 mil HDPE.
Methyl Bromide
79
tarp or saturated with water, a large fraction of the applied MeBr would be expected to escape into the air. Decreases in the total area under a curve in Figure 4 are due to both soil degradation and volatilization. Although the transport of MeBr in soil is largely governed by its intrinsic chemical and physical properties, some soil and meteorological conditions can influence this process by affecting MeBr partitioning over the phases. Of these factors, temperature, soil water content, and content and continuity of soil air spaces are important. With increases in soil temperature, more MeBr is expected to partition into the soil air, the diffusion coefficient is increased, and therefore the movement should be accelerated (Goring 1962). However, little experimental effort has been devoted to measuring MeBr transport in soil, or volatilization from soil, as a function of temperature. Soil water content was shown to be important in controlling MeBr penetration in soil (Kolbezen et al. 1974; Abdalla et al. 1974). Methyl bromide diffuses more rapidly, deeply, and widely in drier soils than in wet soils. Altering soil water content has also been shown to be effective in reducing MeBr volatilization losses from soil. Lower volatilization loss occurred from Greenfield sandy loam with a water content of 18% than from the same soil with 6% water content (Gan et al. 1996). This effect was even more drastic when water was applied to soil surface under tarped conditions (Jin and Jury 1995). Only about 4% of the MeBr was lost after three surface irrigations of 16 mm water in a tarped packed soil column. Soil water content affects MeBr transport by altering the soil void ratio and the continuity of air-filled pores. Theoretically, when a soil is saturated with water, gas-phase diffusion is eliminated and the movement of a volatile chemical will be greatly reduced. Other factors that may affect soil air spaces and distribution include disturbance and compaction of the soil. During mechanized application, void channels are often created by shanks or chisels along the injection lines. It is a common practice during untarped fumigation of MeBr to use a roller behind the shanks to compact the soil surface and to close some of the openings caused by injection. Compaction has been found to result in reduced MeBr diffusion in soil (Knavel et al. 1965). In packed soil columns, MeBr diffused more slowly and less MeBr was lost through volatilization from a soil column packed at 1.7 g cm−3 than from the same soil at 1.40 g cm−3 (Gan et al. 1996). The diffusion of fumigants in soil has been studied using mathematical and numerical simulation techniques since the 1950s. Hemwall (1959) was one of the first to utilize computers to numerically solve a two-dimensional diffusion equation similar to Eq. 11. Two principal goals of this work were the determination of the diffusion pattern and the calculation of the biological control function, also called the concentration–time index: i
B = ∫ C(t) dt
(19)
0
where the value of B gives an indication of the level of biological control. The simulation was used to obtain information on fumigant application methods and the soil MeBr distribution to ensure that efficacious dosages and biological con-
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S.R. Yates, J. Gan, and S.K. Papiernik
trol were achieved. From the simulation, it was found that during the first day the diffusion process was basically radial with rapidly increasing losses to the atmosphere. At later times, the soil concentration was generally uniform with a constantly decreasing flux rate. Since then, numerous studies have been conducted that adopted similar assumptions. Siebering and Leistra (1978) developed a diffusion model to study how moisture, bulk density (e.g., tillage), dosage, and cover time affect control, assessed from concentration–time curves. From the results of their model, they discuss how various soil, chemical, and application factors affect the movement and dissipation of methyl bromide in soils. In a more recent study, Reible (1994) used a one-dimensional diffusion model together with more recent estimates of the model parameters to determine the fate and transport of methyl bromide in soils compared to that of a methyl bromide alternative, methyl isothiocyanate (MITC). Reible suggested that because MeBr is more dense than air, the downward movement of MeBr would occur at a rate of approximately 42 cm/d (assuming 30% porosity and 10−8 cm−2 permeability), but density-driven mass flow was not considered in the model. They also found that a shallow water table would act as a barrier to downward movement and thus enhance volatilization. The simulation included the effects of a film that was assumed to have a constant permeability of 8.31 L/hr/m2. The effect of temperature on the permeability of polyethylene film was not considered, although has been shown to be important (Kolbezen and Abu-El-Haj 1977; Yates et al. 1996c). Additionally, the effect of temperature on the volatilization rate was not considered. Therefore, the estimated flux curves do not show the diurnal fluctuations that are commonly observed in field experiments. Also, the parameters that were used are appropriate for a mean temperature of approximately 20°–25 °C, and the effects of atmospheric resistance above the tarp were not included in the model. The results indicate that approximately 45% of the applied MeBr would be lost to the atmosphere within approximately 14 d when injected at 25 cm and the soil covered with plastic that is removed after 2 d. Several field experiments have shown that the time response to the MeBr flux is highly irregular (Yagi et al. 1993, 1995; Majewski et al. 1995; Yates et al. 1996b) and is caused, in part, by temperature effects on behavior of the plastic (Kolbezen and Abu-El-Haj 1977) and the diffusion coefficient (Reid et al. 1987; Wang et al. 1997b), along with diurnal changes in pressure (Massman and Farrier 1992; Chen et al. 1995) and atmospheric processes (Baker et al. 1996). Sorption of many pesticides does not follow the linear equilibrium assumption, especially when the concentration is constantly changing due to movement in the gas phase. The assumption of a linear equilibrium sorption response is often adopted to simplify the mathematics. Brown and Rolston (1980) showed that the transport of volatile compounds was better described by a first-order kinetic model than by a linear equilibrium assumption. Their experiments were conducted in a laboratory with a steady flow of gas. They found that the retardation factor obtained using the linear equilibrium approach was dependent on the
Methyl Bromide
81
flow rate through the column. The parameters for the first-order kinetic model, however, were not sensitive to flow rate, which is evidence that this model may be more appropriate. To obtain a model suitable to describe field studies, transport in two or three dimensions must be considered. Rolston and Glauz (1982) developed a twodimensional model of MeBr diffusion that included the effects of reversible first-order sorption. They assumed that the input of MeBr into soil was a saturated cylinder of radius r0 and that gas entered the soil after diffusing out of the cylinder. They investigated the use of plastic covers and compared soil gas concentrations measured midway between the injection points to simulated values. The authors found that the simulated values without plastic film matched the field data with film better than the simulation with a film present. This result indicates that either the model did not include important processes, such as heat transport, or one or more of the input parameters were not correctly specified. This represents a challenging test of the model because the data were reported in the literature nearly 10 years earlier; this also points to the difficulty in predicting pesticide transport in field studies. Particularly interesting are the predicted Br− concentrations that show a peak value at the injection depth. Comparing this with field data collected by Yates et al. (1996a) shows that the maximum Br− concentration occurred at the soil surface, which may be caused by higher soil organic matter content in the near-surface soil enhancing the effective degradation coefficient in this region; soil evaporation may also transport Br− to the soil surface.
IX. Assessing Methyl Bromide Volatilization from Soil A. Measurement Methods There are numerous methods for estimating the gas flux from soils to the atmosphere. Three methods used recently to estimate MeBr emissions to the atmosphere include estimating total loss from Br− appearance (i.e., from MeBr degradation, micrometeorological, and enclosure-based methods). 1. Estimating Total MeBr Loss from Br− Appearance The Br−-appearance method assumes that the difference between the MeBr mass applied and mass degraded (i.e., Br− produced) was released into the atmosphere. Therefore, measuring Br− in the soil provides a method for estimating the total atmospheric emission. An advantage of this method is the ease of analyzing the Br− content of soils. A disadvantage is the large number of soil samples necessary to obtain an accurate field-scale estimate of degradation at all depths (Jury 1985; Yates et al. 1996a). Also, no information about the dynamics of MeBr emissions can be obtained using this method. Figure 5 shows the Br− [mg/kg] concentration measured on a dry soil weight basis taken before and after MeBr application during tarped-soil, shallow injection (Fig. 5A) and bare-soil deep injection (Fig. 5B). An estimate of the total
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S.R. Yates, J. Gan, and S.K. Papiernik
Fig. 5. Bromide ion distribution in soil after methyl bromide application at 25 cm and covering the soil surface with high-density polyethylene (HDPE) (A) and after injection into bare soil at 68 cm (B).
MeBr lost to the atmosphere can be obtained from the difference between the initial and final curves and converting from Br− mass to MeBr mass. From Fig. 5A it was estimated that 325 (±164) kg or 39% (±19%) of the applied MeBr was degraded to Br−. Because the MeBr mass (parent material) remaining in the field at the time of Br− sampling was estimated to be less than 0.05% at the time of sampling, the total loss from volatilization is approximately 518 (±164) kg, or 61% (±19%) of the applied MeBr mass. The spatial and measurement variability was found to introduce considerable uncertainty into the Br− mass calculation, as shown by the standard error of ±164 kg, and also produces uncertainty in the estimate of MeBr volatilization. Figure 5B shows the measured soil Br− content before and 3 mon after a deep-injection (68 cm), bare-soil fumigation experiment. From this information it was determined that 78% of the amount applied was detected as Br− (22% volatilized). A second sampling 6 mon after application was conducted to verify the accuracy of the amount degraded and was found to be within 3% of the first sampling, indicating that an accurate field-scale Br− concentration measurement was obtained. The nearly equivalent values for the total mass degraded from each sampling (75%–78%) indicate that sufficient samples were collected to obtain an accurate field-scale average. Field-scale variability has been shown to have considerable effect on the spatial average solute concentration sampled in the field (Jury 1985). This effect has also been shown for MeBr (Yates et al. 1996a) and can be illustrated as follows. If the estimate of the MeBr mass degraded shown in Figure 5A was obtained using 45 background concentrations and 1100 samples taken at the completion of the experiment, the total volatilization loss was estimated to be 298 kg or 35.3%. If, instead, 500 background samples were used, the total volatilization loss was estimated to be 435 kg or 48.4%. When 1100 background samples were used, the MeBr mass loss was estimated to be 518 kg or 61%, which more closely matched the total loss estimates determined by using other
Methyl Bromide
83
methods. This result demonstrates that large errors are possible when solute concentrations are determined in too few samples in a field soil. 2. Chamber Methods The two principal approaches for using chambers to estimate the volatilization rate from soil into the atmosphere are passive (closed system, assumed perfect mixing) and active (or flowing system) chamber methods. The chamber method allows measurement of the volatilization rate over small surface areas compared to other methods, which is both an advantage and a disadvantage. Chambers may offer the only means of measuring volatilization for situations where the source has a small areal extent. For large fields, however, the volatilization rate measured using chambers may be highly variable because of a combination of field-scale spatial variability of properties that affect the mass transfer and the chamber’s small measurement area (Yagi et al. 1993; Yates et al. 1996c; Williams et al. 1999). Each measurement of chemical concentration in the chamber indicates the instantaneous flux. Multiple chambers are required to provide information on the spatial variability of MeBr flux. Flux estimates are sensitive to the placement of the chambers relative to the position of MeBr injection, i.e., distance to the source, and the presence of a chamber can affect the area sampled, especially the local temperature and relative humidity (Clendening 1988). These characteristics of chambers can have a tremendous effect on experimental uncertainty. To characterize the dynamics of MeBr flux and to accurately determine cumulative emissions, numerous samples are required from each chamber over the time period for which volatilization is occurring. Emissions of MeBr from treated fields follow a diurnal variation (Yates et al. 1997), and appropriate sampling intervals must be chosen to adequately represent volatilization throughout the experiment. There have been numerous reviews of the chamber methods, including those by Rolston (1986), Wesely et al. (1989), and Livingston and Hutchinson (1995), as well as articles that describe the use of these methods (Matthias et al. 1980; Hutchinson and Mosier 1981; Reicosky 1990; Harrison et al. 1995) for measuring the surface fluxes of trace gases. The reviews cite several advantages of chambers compared to meteorological methods. For example, chambers are easy to construct, relatively inexpensive to operate and conceptually simple. Chambers can be used to measure gas losses from very small surfaces, which make them suitable for use in laboratory studies to obtain surface emission data to aid in understanding the physical, chemical, and biological aspects of transport and fate of volatile contaminants in soils columns. Passive Chambers Passive or closed chamber systems consist of a container with an open bottom that is placed over the soil surface for a short time period. The concentration of the target gas in the chamber increases with time as the chemical moves from the soil matrix into the chamber. The concentration of MeBr in the chamber is determined at the completion of some time interval (∆t) elapsed since the placement of the chamber on the soil surface. From this
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information and use of various equations, the flux into the chamber can be estimated. A linear model is most commonly used: flux =
V ∆C A ∆t
(20)
where V is the chamber volume and A is the area of soil sampled. This approach assumes complete mixing inside the chamber, constant emissions during the placement period, and no loss mechanism (Matthias et al. 1980; Rolston 1986; Flessa et al. 1995). It has been generally recognized that the simple model in Eq. 20 underestimates the volatilization rate due to a reduction in the gradients as the chemical concentration builds up inside the chamber (Jury et al. 1982; Rolston 1986; Livingston and Hutchinson 1995). To overcome this, more comprehensive nonlinear models have been developed that do not require the constant flux assumption (Hutchinson and Mosier 1981; Livingston and Hutchinson 1995; de Mello and Hines 1994; Valente et al. 1995). For example, Hutchinson and Mosier (1981) use the following: flux =
冋
册
V (C1 − Co)2 C1 − C o ln A ∆t (2 C1 − C2 − Co) C2 − C 1
(21)
where Co is the initial concentration in the chamber. To use Eq. 21, the concentrations C1 and C2 must be measured at two times, ∆t and 2 ∆t. Active Chambers Another simple method for measuring the rate of emission of volatile compounds is the flowthrough flux chamber method (Hollingsworth 1980; Clendening 1988). The flowthrough flux chamber is a closed system device that allows the pesticide emission from a small surface area to be collected. A continuous and uniform flow rate is maintained through the chamber. The flow rate is chosen to avoid MeBr accumulation inside the chamber and to minimize negative pressure inside the chamber, which could cause advective mass flow. Chemical concentrations are determined in the air entering (Cin) and exiting (Cout) the chamber. Once this information is known, the flux is determined using flux =
(flow rate)(Cout − Cin) (sampled area)
(22)
Adsorbent tubes are often used to accumulate MeBr mass from the chamber effluent over some time interval, typically 2–4 hr. Temperature Effects on Chambers One disadvantage of chamber methods is that the pesticide flux measurement is affected by the presence of the chamber covering the sampling area. For example, the temperature and relative humidity have been shown to be affected by the presence of a chamber (Clendening 1988). During a series of experiments, the air temperature inside the chamber was found to be as much as 5 °C higher than the outside air temperature. The
Methyl Bromide
85
relative humidity inside the chamber was found to be from 20% to 50% higher than external values. Using chambers of the same design, Yates et al. (1996c) found that the internal chamber temperature could be as high as 30 °C above the ambient air temperature under certain conditions. A few factors that can alleviate temperature increases inside chambers include using high flow rates and opaque materials whenever possible. Because the permeability of polyethylene plastic to MeBr has been shown to be strongly affected by temperature (Kolbezen and Abu El-Haj 1977; Yates et al. 1997; Papiernik et al. 2002), increased temperature inside a flux chamber can produce biased MeBr volatilization rates. Yates et al. (1996c) demonstrated that a correction should be used to provide a more accurate estimate of the MeBr flux density because of this effect whenever significant heating occurs. In their experiment, uncorrected flux chamber data provided an estimate of ⬃96% of the applied MeBr mass being lost to the atmosphere, nearly 50% greater than the total loss estimates from micrometeorological and the appearance of Br− methods. After correcting for temperature increases inside the active chambers, estimated total MeBr emission was approximately 59%, about 5% less than estimated from micrometeorological and appearance of Br− methods. This finding indicates the importance of designing chambers that minimize internal heating. 3. Micrometeorological Methods The meteorological techniques require that emissions occur over a large surface area, so that relatively stable gas concentration profiles can be established and detected above the soil surface. Reviews and comparisons of various meteorological techniques are available (Fowler and Duyzer 1989; Wesely et al. 1989; Majewski et al. 1990; Denmead and Raupach 1993). Methods for measuring the volatilization rate using micrometeorological information are fairly complex, require numerous measurements of MeBr concentration and other atmospheric parameters, and may require assumptions concerning the behavior of the atmosphere. Advantages are that the methods are well tested, they provide a field-scale average total emission rate, and they provide information on the dynamics of the volatilization process. Aerodynamic Method The aerodynamic method is based on atmospheric gradients of wind speed, temperature, and concentration and provides a measurement of the pesticide flux from the soil surface (Parmele et al. 1972; Brutsaert 1982; Majewski et al. 1989). The method requires a spatially uniform source and a relatively large upwind fetch so that the atmospheric gradients are fully developed. The fetch requirements are generally assumed to be from 50 to 100 times the height of the instruments, which for large agricultural fields is typically greater than 0.5 m in height. Chemical concentrations, wind speed, and temperature are determined at multiple heights at the sampling location to define the gradient. To characterize the temporal variability in flux, samples are collected over relatively short time intervals, typically 2–4 hr. Adsorbent tubes are often used to accumulate the volatilized mass throughout the sampling interval.
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S.R. Yates, J. Gan, and S.K. Papiernik
The aerodynamic method was originally developed for use under neutral atmospheric conditions. Using empirical relations, however, the method can be extended to stable and unstable atmospheric conditions. Numerous stability corrections have been proposed (Fleagle and Businger 1980; Brutsaert 1982; Rosenberg et al. 1983). The aerodynamic equation, suitable for general atmospheric stability conditions, is: fz(0, t) = k2
[c¯ 1(t) − c¯ 2(t)][u¯ 2(t) − u¯ 1(t)] ϕm(t) ϕc(t) ln(z2/z1)2
(23)
where k ⬇ 0.4 is von Ka´rma´n’s constant, fz(0, t) is the interval-averaged vertical flux at the soil surface, u¯(t) is the interval-averaged wind speed [m/s], z is height above the soil surface [m] (note: z2 > z1), and c¯1(t) is the interval-averaged concentration [µg/m3] at height z1 above the soil surface. The gradient-based stability corrections for a particular time interval, t, can be written as (Rosenberg et al. 1983): 1
ϕc = ϕm = (1 − 16Ri)− 3
Ri < 0
unstable
1 3
Ri > 0
stable
ϕc = ϕm = (1 + 16Ri)
+
(24)
where it is assumed that the stability functions for momentum and the concentration are the same. Majewski et al. (1995) used a slightly different equation to account for atmospheric stability, and several other stability corrections have been proposed (Fleagle and Businger 1980; Brutsaert 1982; Rosenberg et al. 1983). The Richardson number, Ri, is defined as Ri =
冋册
g ∂T ∂u −2 T ∂z ∂z
(25)
where g is the gravitational acceleration (i.e., 9.8 m/s2), and T(t) is the absolute temperature [K]. The gradient Richardson’s number is one means for characterizing the importance of buoyancy and mechanical mixing on the turbulence. Theoretical Profile Shape Method The theoretical profile shape method (Wilson et al. 1982) can be used to determine the volatilization rate from field experiments conducted on a circular plot. This method has advantages over the aerodynamic method in that (1) the large fetch requirement is not necessary, (2) measurements of the air concentration and wind speed are needed at only one height, and (3) the sensor is placed at a height that is relatively insensitive to the atmospheric stability so temperature and wind gradients and stability corrections are unnecessary. This approach is based on the trajectory simulation model described by Wilson et al. (1981a–c). Wilson et al. (1983) and Majewski et al. (1990) have used this method, among others, to determine the rate of pesticide and ammonia volatilization from field experiments. Yates et al. (1996b, 1997) adapted the method so that MeBr volatilization from rectangular fields could be estimated.
Methyl Bromide
87
The flux density is estimated from [u¯ (t) c¯ (t)] *Zinst (26) Ω where interval-average values of the wind speed, u¯(t), and air concentration, c¯(t), are obtained at the instrument height, Zinst. Flux can be obtained by determining the ratio of the horizontal to vertical flux, Ω, using the trajectory simulation model discussed below. This ratio depends on surface roughness and upwind fetch distance (i.e., the radius of the circular plot) but does not depend on wind speed. fz(0, t) =
Trajectory Simulation The theoretical profile shape method is based on the trajectory simulation model. This model is used to simulate pesticide transport in the atmosphere using a particle-tracking algorithm and assuming that the atmosphere experiences conditions of inhomogeneous turbulence. Wilson et al. (1981a–c) give a complete description of the method. The simulation of a particle of air mass proceeds in a hypothetical atmosphere in which the Eulerian velocity, σw(z), time [τ(z)] and length [Λ(z) = σw(z)τ(z)] scales are assumed to vary only in the vertical direction. It is further assumed that the horizontal wind speed varies only in the z direction, the timeaveraged value of the vertical wind speed is zero, the pesticide source is spatially uniform, there are no sources of pesticide outside the treated area, no degradation of pesticide occurs once it is in the atmosphere, and the surface roughness of the field, zo, is assumed to be constant. A source area extending from 0 ≤ x ≤ Xmax is discretized into M sections of equal length. A large number of particles is emitted from each section and tracked until the particles reach the collector (located at Xmax) which coincides with the position of the sampling mast in the field. During the simulation, the instantaneous vertical position of each particle is obtained and used to determine the current horizontal position increment for the current time step. Once a particle has reached the collector, the count of the appropriate height level, Z, is incremented. As the number of particles released increases, the statistical character of the vertical distribution becomes fixed and the ratio of horizontal to vertical flux, Ω, as well as the instrument height, Zinst, becomes known. To obtain Ω, the simulation is conducted for strongly stable, strongly unstable and neutral atmospheric conditions. Plotting the results against height produces a curve similar to that shown in Fig. 6. The height where the three curves more or less converge is the height where the sensor is positioned. At this location, the effects of atmospheric stability are minimized. One difficulty using this method is the determination of the instrument height before initiating the experiment; this involves estimating surface roughness, which may not be known until after the experiment begins (i.e., when the plastic is placed on the field, at the time of application). Because the upwind source distance varies with the wind direction for rectangular fields, the trajectory simulation must be conducted for several upwind source distances ranging from the smallest to the largest distance between the
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S.R. Yates, J. Gan, and S.K. Papiernik
Fig. 6. Using the trajectory simulation model (Wilson et al. 1982), a height in the atmosphere can be found that is relatively insensitive to atmospheric stability. For this example, the instrument height is approximately 288 cm and Ω is approximately 15.
sampling mast and edge of the field. Then, a relationship can be developed between wind direction, instrument height, and Ω, which may be used to estimate the flux using the average wind direction for the sampling interval. Integrated Horizontal Flux Method The integrated horizontal flux method (Denmead et al. 1977; Wilson et al. 1982; Majewski et al. 1990) can be used to estimate the surface flux when the concentration, c(z), and horizontal wind speeds, u(z), in the atmosphere are known as a function of height. Measurements of chemical concentration and wind speed gradients are similar to those used for the aerodynamic method. Assuming a spatially uniform source, the flux is estimated from a statement of mass balance, that is: fz(0, t) =
1 L
∞
∫
u¯ (z) c¯ (z) dz
(27)
0
From this equation, the mass emitted from the surface upwind from a sampling point is equal to the mass that passes through a vertical plane of sufficient height to capture all the mass (i.e., of infinite extent) located at the sampling point. To use this method, the concentration profile at several heights must be determined, and the distance of the source area upwind from the sampling mast must be known. An advantage of this method over the aerodynamic method is that cor-
Methyl Bromide
89
rections for atmospheric stability are not needed because this approach is based on principles of mass balance. B. Field Experiments to Determine Methyl Bromide Volatilization Since the mid-1990s, several experiments have been conducted to obtain information on MeBr emissions from typical agricultural operations. The results from these studies are summarized in Table 5. Various methods were used to estimate the emission rate, including an increase in soil Br− concentration as a result of MeBr degradation (Yates et al. 1996a), the atmospheric flux method (Majewski et al. 1995; Yates et al. 1996b), and the enclosed flux chamber method (Yagi et al. 1993, 1995; Yates et al. 1996c). Every method has advantages and disadvantages that often make the interpretation of the experimental results somewhat difficult. However, for determining the total emission, all the methods should provide reasonably accurate results. 1. Yagi et al. (1993) Yagi et al. (1993) conducted an experiment in Irvine, California, to measure the MeBr emission from a fumigated southern California field using four passive flux chambers. MeBr was applied at a depth of approximately 25 cm and the soil surface was covered with low-density polyethylene plastic film. The authors originally estimated that 87% of the total MeBr applied to the field escaped into the atmosphere. This estimate was revised to 74% ± 5% (Williams et al. 1999) by eliminating the data from a chamber that covered tarp material with a hole. The estimates of MeBr emissions measured during this study are the highest reported for MeBr injection at shallow depth and the soil surface covered with plastic. The high emission rates are probably due to a combination of factors such as use of low-density polyethylene plastic, which is permeable to MeBr vapors (Kolbezen and Abu-El-Haj 1977), the presence of a high bulk density and moist soil layer at 60 cm depth. This value is also higher than expected given other estimates based on mathematical models (Albritton and Watson 1992; Singh and Kanakidou 1993), but was similar in magnitude to the losses observed in glasshouse studies (de Heer et al. 1983). To verify these results, the authors returned to the field to collect Br− information to provide mass balance information (Yagi et al. 1995). 2. Yagi et al. (1995) The investigators conducted a second experiment in a nearby field using the same procedures as their first experiment (Yagi et al. 1993). For this experiment, high-density polyethylene (HDPE) plastic was used to cover the field and five flux chambers were used to measure emissions. They found that only 34% of the applied MeBr escaped to the atmosphere. This value is more than 50% lower than the result of their first experiment, which included a low-density polyethylene tarp. Variability in the emission measurement is expected for several reasons: (1) only 10–15 samples of the volatilization rate were obtained during each 7-d experiment, generally at the high point during the day; (2) only a few soil samples were taken to measure Br− concentrations,
90
Table 5. Total amount of methyl bromide volatilized during the experiment and mass balance.
Flux chamber Flux chamber Aerodynamic, profile Aerodynamic, profile Br− appearance Aerodynamic, discrete Aerodynamic, profile Theoretical profile shape Integrated horizontal flux Flux chamber Br− Appearance Aerodynamic, discrete Aerodynamic, profile Theoretical profile shape Integrated horizontal flux Flux chamber Flux chamber (location 1) Flux chamber (location 1) Flux chamber (location 2)
Surface cover
25–30 35 25–30 25–30 25–30
PE PE Bare PE HDPE
25–30
HDPE
69
25–30
Bare
LDPE
Number of sampling intervals 14 14 31 33* 2 105 105 105 105 107 2 164 164 164 164 173 12 17 14
MeBr Percent Number of Field mass MeBr measurements Total number of size applied volatilized per interval measurements (ha) (kg/ha) (%) 4 chambers 5 chambers 6 heights 6 heights 1100 samples 2 heights 6 heights 6 heights 6 heights 3 chambers 1050 samples 2 heights 9 heights 9 heights 9 heights 4 chambers 3 chambersa 8 chambers 5 chambersb
56 70 186 198 2200 210 630 630 630 321 2098 328 1476 1476 1476 692 36 136 70
17 1 6 4 3.5
256 243 199 262 240
3.5
240
3.5
322
17 17 1
256 208 243
Mass balance (%)
87 — 34 104 32 — 89 — 61 — 62 101† 67 106† 60 99† 70 108† 59 97† 21 ± 2.6 — 4.5 ± 1.3 83† 3.1 ± 0.1 82 2 81 1.9 ± 0.2 81 4.9 ± 3.1 84 74 ± 5a 97 ± 5a 63 ± 12 94 ± 12 36 ± 6b 106 ± 11b
Reference Yagi et al. (1993) Yagi et al. (1995) Majewski et al. (1995) Yates et al. (1996a)
Yates et al. (1997)
Williams et al. (1999)
S.R. Yates, J. Gan, and S.K. Papiernik
Method used
Injection depth (cm)
Table 5. (Continued).
Method used
Injection depth (cm)
Surface cover
35
HDPE
25
Flux chamber (10-d cover period)
25
Flux chamber (15-d cover period)
25
Br− Appearance
25
Br− Appearance
60
HDPE Hytibar Hytibar HDPE Hytibar Hytibar HDPE Hytibar Hytibar Bare HDPE HDPE + water Bare HDPE Hytibar
16 17 13 114 162 114 114 162 114 114 162 114 2 2 2 2 2 2
MeBr Percent Number of Field mass MeBr measurements Total number of size applied volatilized per interval measurements (ha) (kg/ha) (%) 7 chambers 8 chambers 7 chambers 2 2 2 2 2 2 2 2 2 75 75 75 75 75 75
112 136 91 228 324 228 228 324 228 228 324 228 150 150 150 150 150 150
1 1 — 0.002c
0.002c
0.002c
0.002c
0.002c
233 257 310 280 210 140 280 210 140 280 210 140 345 241 363 427 368 508
24 ± 5 45 ± 8 50 ± 9 68 36 39 56 2 3 67 3 1 87 59 <42 60 15 <15
Mass balance (%)
Reference
98 ± 9 86 ± 15 97 ± 13 95d Wang et al. (1997a) 105e 100d 87d 102e 98d 108e 99d 95d — Wang et al. (1997c)
Methyl Bromide
Flux chamber (location 2) Flux chamber (location 2) Flux chamber (location 3) Flux chamber (5-d cover period)
Number of sampling intervals
—
91
HDPE: high-density polyethylene; LDPE: low-density polyethylene; PE: polyethylene. †Mass applied 843 kg; measured mass remaining 0.26 kg (from Yates et al. 1996a); §Values in parentheses are standard deviations; *In addition, 12 samples were estimated. aReanalyzed data from Yagi et al. (1993). bReanalyzed data from Yagi et al. (1995). cPlot size 3.4 m × 4.9 m. dAfter 15 d; eAfter 21 d.
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which have been shown to be highly variable (Jury 1985; Yates et al. 1996a); and (3) degradation of soil MeBr is highly dependent on actual soil conditions. An additional source of variability may be the internal chamber temperature, which has been shown to affect HDPE permeability. Yagi et al. (1993, 1995) did not correct their volatilization rates for this effect. The estimated 34% loss rate is within the 30%–60% expected loss range noted in the Montreal protocol. 3. Majewski et al. (1995) Two field experiments were conducted in Monterey County, California, from Oct. 26 to Nov. 4, 1992. The fields (Salinas clay loam) were separated by a distance of approximately 6 km. MeBr was injected at a depth of 25–30 cm, and one field was covered with a standard high-density polyethylene plastic film, with the other left uncovered. The application rate for the tarped experiment was 392 kg/ha and for the bare soil experiment was 203 kg/ha. In both experiments the flux density was measured using the aerodynamic method (Parmele et al. 1972). The aerodynamic method produces a large-scale average volatilization rate and is relatively insensitive to small-scale variability that may occur when using chambers. Although an error analysis was conducted, insufficient information was obtained for a mass balance; therefore, there was no independent measure of the total emission. Majewski et al. (1995) found that 32% of the applied MeBr was emitted into the atmosphere from the tarped field during the first 9 d following application. This value is approximately the same as that from the second study of Yagi et al. (1995) and falls into the 30%–60% range noted in the Montreal protocol (Albritton and Watson 1992). For the bare soil experiment, approximately 89% of the applied fumigant was lost via volatilization. 4. Yates et al. (1996a–c) Yates et al. (1996a–c) conducted an experiment at the University of California’s Moreno Valley Field Station on a 4-ha field during August and September 1993. The soil in this field is a Greenfield sandy loam. MeBr (99.5% purity) was applied at a depth of 25 cm, at a rate of 240 kg/ha, and the field was covered with 1-mil high density polyethylene plastic. Estimates of the MeBr emission rate and total loss were obtained using flux chambers, micrometeorological methods, and by estimating total loss from Br− appearance. Using the micrometeorological methods (e.g., aerodynamic, theoretical profile, and integrated horizontal flux methods), total emission was estimated to be 62%–70% (±11%) of that applied. Data from the flux chambers gave a total emission loss of about 59% of the applied mass and is from 3% to 10% lower than the estimates from the micrometeorological methods. Cumulative emissions based on Br− appearance totaled 61% of the applied MeBr. A mass balance was calculated for each method used to estimate the flux (see Table 5). The average mass recovery using all flux methods was 103% (±10%) of the applied mass. The range in the mass balance percent (i.e., percent of applied mass that was measured) was 97%–108%. The estimated 60% loss is at the high end of the range noted in the Montreal protocol and by Reible (1994). This experiment
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was conducted under warm, dry conditions using multiple methods for measuring the volatilization rate. Because all methods produce supporting estimates, it is likely that 60% total loss is correct and that large fractions of applied MeBr are lost when fumigation is performed under these soil and environmental conditions. The fraction of the applied MeBr mass volatilized in the experiment of Yates et al. (1996a–c) is approximately double the value reported by Majewski et al. (1995), who estimated the total loss to be approximately 32%, which probably results from regional differences in the climatic and soil conditions between the central coast and inland southern California. Lower temperatures in Monterey would cause a reduction in the diffusion through polyethylene plastic (Kolbezen and Abu-El-Haj 1977) and increase the soil residence time; this would facilitate greater MeBr degradation in the soil and reduce total loss to the atmosphere. The range for total emissions described in Yates et al. (1996a–c) also differs from the results of Yagi et al. (1993, 1995), who reported values of approximately 87% and 34%, respectively, for experiments with similar application methodology. 5. Yates et al. (1997) MeBr volatilization rate was determined in an 4-ha agricultural field after injection at 68 cm; results were compared to an earlier experiment where MeBr was injected at 25 cm and the surface covered with high-density polyethylene plastic (Yates et al. 1996a–c). Three independent methods were used to estimate the total MeBr volatilized after application, i.e., the appearance of soil Br−, the flux chamber, and micrometeorological methods. When injected deep in soils, the MeBr volatilization rate continued at relatively high values for more than 7 d after application. It was observed that the total MeBr mass emitted from the field was significantly less than the earlier experiment, which was attributed to deep injection, cooler air temperatures, and smaller thermal gradients. Total emissions estimate obtained from Br− content sampling was 21% of the applied MeBr. Estimates obtained from direct flux measurements ranged from 1.9% to 4.9%. Mass recovery ranged from 81% to 84% of the applied mass, with an average of 82%. Comparison of the direct methods for measuring volatilization rate with the estimate of total emissions from MeBr degradation suggests that for deep injection using only two shanks, the initially high MeBr gas pressure may cause localized evaporation to play a significant role in the volatilization process. This process needs to be further studied to develop methods for controlling volatile losses. 6. Williams et al. (1999) Emission studies of Yagi et al. (1993, 1995) and four additional experiments were summarized. Three field sites near Irvine, California, were used. Two experiments were conducted at site I where total emissions were 74% ± 5% (Yagi et al. 1993) and 63% ± 12%. Mass balance for this site was between 94% and 97%. At a second site, three experiments were conducted. Total emissions were somewhat variable with measurements of 36% ± 6%, 24% ± 5%, and 45% ± 8%, respectively, in 1993, 1994, and 1995 experiments. Mass
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balances were from 86% ± 15% to 106% ± 11%. A sixth experiment conducted at a third site yielded 50% ± 9% total emissions with a 97% ± 13% mass balance. The emission rate was highly dependent on the thickness of the plastic film used to cover the soil surface during fumigation. The effect of soil carbon content, nitrogen content, and pH on emissions was also investigated. 7. Wang et al. (1997a,c) Smaller-scale outdoor studies (plot size, ⬃17 m2) were conducted to determine MeBr emissions in untarped and tarped plots. Cumulative emissions were ⲏ60% for bare plots and plots tarped with high-density polyethylene when MeBr was injected at 25 cm. A large decrease in emissions was observed with the use of a low-permeability tarp (Hytibar), with volatilization losses <5% of the applied MeBr when the tarp remained in place for at least 10 d.
X. Potential Methods for Minimizing Volatilization Many soil chemical processes affect the fate and transport of soil fumigants, including MeBr. Containment, degradation, and soil gas concentration (i.e., effective dosage) must be controlled to reduce emissions while maintaining adequate pest control. Unless each of these factors is controlled, unacceptable emissions or inadequate pest control will likely occur. Two important factors that may affect the transformation and distribution of MeBr and its ultimate volatilization into the air are depth of placement and the use of soil surface cover. These two factors have been frequently used to alter MeBr distribution to reach adequate control in the specified target zones. Depending on the target pest, MeBr can be applied at the soil surface under plastic film to as deep as 60–100 cm below the soil surface. Deeper placement consistently results in deeper penetration. Abdalla et al. (1974) found that application at 76–81 cm without a soil cover resulted in gas distribution at concentrations sufficient for nematode kill as deep as 244 cm. Kolbezen et al. (1974) detected adequate dosages at 300–360 cm when MeBr was applied at 90 cm. Although these early studies were mostly designed for achieving better nematode control in deep soil layers, they demonstrate that MeBr downward diffusion is encouraged by deep application. More recent research has been directed at investigating other methods for controlling emissions. A. Containment Containment is necessary to hold the gas at the treatment location and provide sufficient fumigant concentrations and time for pest control. Without adequate containment, a large fraction of the applied MeBr will be lost to the atmosphere. Containment is difficult because of the high vapor pressure (approximately 1420 mmHg at 20 °C) and low boiling point (3.56 °C) of MeBr, which result in a large fraction of MeBr existing in the vapor phase at temperatures and pressures that normally occur in the field. Movement in the vapor phase occurs more
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rapidly than in the liquid phase, in part because the diffusion coefficient in the gas phase is larger than in the liquid phase. Therefore, pesticides that have a high vapor pressure move easily through soil (Goring 1962; Kolbezen et al. 1974; Reible 1994). Factors that affect containment include use of plastic, the properties of plastic, injection depth, soil bulk density, soil water content, soil cracking, and other mechanisms that promote or retard movement. For example, shortly after injection, MeBr movement may be dominated by pressure-driven flow in response to phase-change expansion and the initially high gradients near the injection point; this can cause MeBr to quickly move to the soil surface where it can escape into the atmosphere. Other processes may also be important in moving fumigants through the root zone. For example, changes in barometric pressure (Massmann and Farrier 1992), pressure effects caused by wind at the surface, and density sinking (Goring 1962) all may induce a mass flow. Although it may be possible to take advantage of many soil factors to aid in containing MeBr, the inherent spatial variability of soils makes it difficult to ensure emissions control for every situation. 1. Plastic Film Barriers A commonly used and predictable method to improve containment and reduce the amount of MeBr leaving the treated soil is the use of plastic films. Covering the soil with plastic film, or tarping, was introduced early in the development of soil fumigation with MeBr to enhance efficiency and reduce toxicity to workers and residents. Covering the field with plastic can reduce the amount of MeBr volatilized by inhibiting transport from the soil into the atmosphere. In early studies, it was demonstrated that tarping encourages longer retention of MeBr in soil. Kolbezen et al. (1974) reported increased concentrations in surface soil under tarped conditions, but they later noticed that polyethylene film was a poor cover material compared to other plastics in retaining MeBr in the soil (Kolbezen and Abu-El-Haj 1977). The ineffectiveness of polyethylene tarps in preventing volatilization was realized later by many other workers (de Heer et al. 1983; Rolston and Glauz 1982; Jin and Jury 1995). This ineffectiveness is caused not only by the high permeability of the polyethylene film to MeBr but also by the short cover time normally used in agricultural fumigations. Advantages of using films are that their properties and condition are known in advance and films are more uniform in space and time compared to soil. Therefore, there may be a higher certainty of effective containment when films are used compared to soil water-based methods. Also, the level of containment can be controlled by altering the plastic material used. New plastics are available that are very impermeable to MeBr diffusion, which has led to the need for methods to estimate film permeability. Kolbezen and Abu-El-Haj (1977) described a steady-state method to measure the permeability of MeBr to various agricultural films. They found that permeability to MeBr at 23 °C was 8.2 mL hr−1 m−2 for low-density polyethylene and 1.4–2.7 mL hr−1 m−2 for high-density polyethylene film. They also investigated the effect of other plastic material and the effect of temperature on film permeability.
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Permeability of plastic films to fumigant vapors has been determined by measuring the flux of fumigant through the film under steady-state conditions (Kolbezen and Abu-El-Haj 1977), which produces a permeability measurement that is dependent on the concentration gradient maintained across the film. A new approach for estimating film permeability by measuring the mass transfer coefficient (h) of fumigant compounds across agricultural films was presented by Papiernik et al. (2001). The h is a measure of the resistance to diffusion, which, unlike other measures of permeability, is a property of the film–chemical combination and independent of the concentration gradient across the film. This method uses static sealed cells; fumigant vapor is spiked to one side of the film and the concentrations on both sides of the film are monitored until equilibrium. An analytical model is fitted to the data to obtain h. This model relies on a mass balance approach, and includes sorption to and diffusion across the film membrane. This approach was used to determine the h of MeBr and alternative fumigants across plastic films (Table 6) and to assess the impact of environmental conditions on the permeability of 1-mil HDPE to fumigant vapors. Traditional 1-mil (i.e., 0.025 mm) high-density polyethylene is relatively permeable to MeBr (Table 6) (Kolbezen and Abu-El-Haj 1977; Chitwood and Deshusses 2001; Papiernik et al., 2001). Permeability of 1-mil HDPE is affected by ambient temperature. For MeBr, h increased exponentially with increasing temperature, following the trend h = 0.1319 e(0.0505 T) (Fig. 7), corresponding to an increase in h by a factor of 1.7 per 10 °C increase in temperature from 20° to 40 °C. This increase in permeability with temperature has the potential to have a large impact on emissions. Diurnal variations in MeBr flux density from the soil surface are observed in field studies, where flux density is much higher during the day than at night (Yates et al. 1996b). Diurnal variation has been attributed primarily to a higher rate of MeBr diffusion through the plastic at higher temperatures. Standard 1-mil HDPE is more permeable to MeBr alternatives than to MeBr (see Table 6), indicating the importance of alternative emissions reduction strategies for these compounds. Low-permeability films have been developed that Table 6. Mass transfer coefficient (cm hr−1) of fumigants across plastic films at 20 °C. Fumigant
1-mil HDPE Aa
1-mil HDPE Ba
4-mil black HDPE
Silver mylarb
Methyl bromide Propargyl bromide cis-1,3-D trans-1,3-D Chloropicrin
0.37 ± 0.02 1.50 ± 0.07 2.0 ± 0.2 3.7 ± 0.3 0.62 ± 0.07
0.63 ± 0.04 2.64 ± 0.2 3.7 ± 0.4 5.6 ± 0.8 1.11 ± 0.09
0.14 ± 0.01 0.48 ± 0.02 0.56 ± 0.05 1.35 ± 0.4 0.25 ± 0.03
<4.4 × 10−6 <1.3 × 10−5 <2.5 × 10−5 <4.3 × 10−5 <2.3 × 10−4
aPermeation
of fumigants through two different HDPE films of 1.0-mil thickness was tested. permeation of any compound observed through silver mylar was detected; h is based on the detection limit of each compound (Papiernik et al. 2001). bNo
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Fig. 7. Temperature dependence of the mass transfer coefficient (h) for HDPE film and MeBr.
have permeability hundreds to possibly thousands of times lower than that of 1mil HDPE. Environmental conditions other than temperature do not appear to have a large impact on the permeability of 1-mil HDPE to fumigant vapors. Fumigant compounds are often applied to soil in mixture; for example, chloropicrin is added to MeBr formulations. The mass transfer coefficients of MeBr, chloropicrin, and other fumigants across 1-mil HDPE are the same when the compounds are applied together as when they are applied alone (Papiernik and Yates 2002). The permeability of a UV-stabilized HDPE did not change significantly during a period of field exposure typical of conditions of MeBr fumigation in California (Papiernik and Yates 2002). During field fumigation, water condenses on the underside of the film surface. Papiernik and Yates (2002) found that the presence of a water film, even a continuous layer several millimeters in thickness, did not significantly impede diffusion of MeBr or other fumigant compounds through 1-mil HDPE. The rate of diffusion across HDPE film with condensed water appears to be limited by the rate of diffusion of fumigant compounds across the film, with the water
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having only a minimal effect of a short delay, of the order of minutes, in the initial appearance of fumigant at the film surface because of the time required for dissolution into and diffusion through the water phase. Although the presence of condensed water may not impact tarp permeability, studies have shown that increasing the soil water content under soil tarps decreases cumulative emissions of soil fumigant compounds. Slower transport from the application depth to the soil surface in soils with high moisture allows increased time for degradation in the soil. In a soil column study where the soil surface was tarped with PE and the column was exposed to diurnal temperature variations, soil water accumulated at the soil surface. This accumulation resulted from condensed water on the tarp being redeposited on the soil surface and from diurnal heat variation resulting in upward flux of water vapor at night (Jury et al. 1996). Increased water content at the soil surface had a large impact on cumulative emissions in PE-tarped columns, so that the application of water under 1-mil HDPE was much more effective at reducing cumulative MeBr emissions than was surface sealing with 1-mil HDPE alone (Jin and Jury 1995). Because the results of film permeability studies indicate that condensed water on the film does not significantly reduce the apparent permeability of HDPE (Papiernik and Yates 2002), the effect of application of irrigation water on reducing emissions is likely due to the increased water content decreasing the rate of fumigant diffusion in the soil, so that the flux of fumigant from the soil surface to the film is greatly reduced in soils with high water content. Using model simulations, Reible (1994) showed that even though standard polyethylene film is somewhat permeable to MeBr, it provides an effective, temporary barrier. Following a 25-cm injection, extending the cover time from 2 to 7 d reduced the volatilization loss from 53% to 33%. The total volatilization loss would depend heavily on the selected degradation rate. In packed columns, covering the soil surface with 0.025-mm high-density polyethylene film for 2–3 wk generally reduced volatilization loss by 40% (Gan et al. 1997a). Less-permeable plastics were extensively tested on various scales (Kolbezen and Abu-ElHaj 1977; de Heer et al. 1983; Gamliel et al. 1997; Chakrabarti et al. 1995; Daponte 1995; Wang et al. 1997a, 1998; Papiernik et al. 2001), and their usefulness in reducing MeBr volatilization has attracted attention during the past few years. However, most of these improved plastic tarps are considerably more expensive than polyethylene, and their use may be practical only if the reduction in application rates from their use can compensate some of the cost differences. A successful example is the substitution of low-density polyethylene film with Saranex in the Netherlands in 1980s (Hamaker et al. 1983). By the use of this less-permeable material and extended cover time from 2 to 10 d, the application dosage was reduced by half and, subsequently, less volatilization of MeBr was reported (Hamaker et al. 1983; Wegman et al. 1983). Wang et al. (1997a) also demonstrated a large reduction in MeBr volatilization using reduced application rates and high-barrier plastic films. In plots with 50% of the standard MeBr application rate (140 kg/ha) and Hytibar as a surface tarp, efficacy against citrus nematodes (Tylenchulus semipenetrans) and yellow nutsedge (Cyperus esculen-
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tus) seeds was as good as in the standard fumigation (280 kg/ha) using polyethylene tarp, but efficacy against rhizoctonia fungi (Rhizoctonia solani) was not adequate (Wang et al. 1997a). Efficacy against R. solani improved with application of 210 kg/ha MeBr (75% of the standard application rate) and Hytibar tarp. In the Hytibar-tarped plots, <4% of the applied MeBr was volatilized when the tarp remained on the soil surface for at least 10 d; in polyethylenetarped plots, MeBr emissions were greater than 50% (Wang et al. 1997a). Table 7 provides a summary of the total emission in percent of applied MeBr for both tarped and untarped treatments following injection into soil columns (Gan et al. 1997a). These results indicate the importance of injection depth and use of a plastic tarp in reducing MeBr volatilization. When the soil surface was not covered with polyethylene, MeBr volatilization was extremely rapid, with as much as 80%–90% of the total loss occurring during the first 24 hr. In contrast, when a tarp was present, maximum volatilization flux was significantly smaller, with only 30%–45% of the overall loss occurring during the first 24 hr. Although measurable volatilization rates continued for a longer time (7–10 d) compared to the untarped columns (3–4 d), total emissions were significantly lower in tarped columns (Table 7). Similar results were observed in two parallel field experiments (Majewski et al. 1995). In an untarped field, MeBr emission after injection at 25–30 cm was 89% over the first 5 d after application, while in a tarped field located 6 km away the emission rate was 32% over the first 9 d. Based on these results and the few recently reported field studies, it is clear that the emission rate is reduced in a tarped field compared to untarped conditions when MeBr is injected at shallow depth (20–30 cm). Films with lower permeability should produce even greater emission reductions. Also, because MeBr is retained in the soil much longer under films with lower permeability, it should be possible to reduce the application rate without sacrificing the fumigation efficacy (Hamaker
Table 7. Effects of injection depth and use of plastic films.
Injection depth (cm) Tarped columns 20 30 60 Nontarped columns 20 30 60 aTo
Total emissions (measured) (%)
Total degradation (%)
Mass balance (%)
Total emissions corrected using diffusion modela (%)
59 52 45
36 39 46
94 91 91
43 37 26
91 83 60
12 15 36
102 98 96
82 71 38
correct for the presence of a lower impermeable boundary, see section VII.C.1.
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et al. 1983; Wang et al. 1997a). Reducing the application rate when high barrier films are used may provide a means for producing significant decreases in emissions. Table 8 shows some recent estimates of total MeBr emissions for soils left bare or covered with plastic film after application. It is clear that the use of HDPE has a beneficial effect on reducing MeBr emissions and that using virtually impermeable films appears to hold great promise for reducing emissions to near-zero levels. 2. Application Depth Depth of MeBr application is also an important factor affecting the amount escaping into the atmosphere. In laboratory soil columns, when the application depth was increased from 20 to 60 cm, MeBr emission rates decreased by 54% under untarped conditions and 40% under tarped conditions (Table 7). Combining deep injection with the use of a surface tarp has the potential to significantly reduce MeBr emissions (see Table 7). This finding supports the results from a field experiment (Yates et al. 1997) in which 21% of the applied mass was volatilized when MeBr was injected deep in the soil (68 cm) and the soil surface left uncovered; this is 66% less than the total emissions rate (62%) observed in an adjacent field for which shallow injection and a high-density polyethylene tarp were used (Yates et al. 1996a–c). Lower emissions may have been attributable to deeper injection depth and a cooler average air temperature. These results agree with the predictions made by Reible (1994) using a vapor transport model. Under hypothetical conditions, it was estimated that increasing the injection depth from 15 to 45 cm would decrease the MeBr emission rates from 53% to 28% when the soil was tarped. From these findings, it can be concluded that placing MeBr at a greater depth is another effective approach for minimizing its emission into the air during soil fumigation. The combined effect of injection depth and surface cover on total MeBr emissions is shown in Fig. 8, which assumes that a tarp covered the field for 5, Table 8. Effects of film presence and type on total emissions for shallow injection (20–35 cm). Laboratory columnsa Surface barrier Bare HDPE Hytibar aGan
Field plot experimentsb
Field experimentsc
Cover Total Cover Total Cover Total period (d) emission (%) period (d) emission (%) period (d) emission (%) na 8 8
44–90 37–83 2e
na 15 15
87 67 <5
na 4,5,8 na
8 32–67d na
et al. (1996, 1997a, 1998d, 2000b); bConducted in small field plots (Wang et al. 1997a); cYagi et al. (1993, 1995); Majewski et al. (1995); Yates et al. (1996b,c); Williams et al. (1999); dRemoved highest and lowest value from available data; eCumulative emissions at 8 days. Total emissions including losses after the tarp was removed were 34% less than HDPE.
Methyl Bromide
Fig. 8. The effect of injection depth, cover period, and surface cover on total emissions of methyl bromide.
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10, or 30 d after application. Conditions used for the simulation were soil diffusion (Ds = 1450 cm2 d−1), soil degradation (µ = 0.06 d−1) (e.g., half-life of 12 d), temperature 30 °C, and a mass transfer coefficient for HDPE (h) of 18 cm d−1. The mass transfer coefficient for a typical low-permeability film (Hytibar) was assumed to be 200 times less than that of HDPE, and a large value for h was used for bare soil conditions (i.e., h = 1000 cm/d). For typical application depths (i.e., 25–75 cm), it is clear that leaving the soil uncovered will result in large emissions to the atmosphere (e.g., 87% for 20-cm injection depth). Using a HDPE film for 5 d reduces emission by about 20% compared to bare soil, but increasing the cover time to 10 d offers little improvement in reducing emissions. For shallow injection, a substantial reduction in emissions can be obtained using low-permeability films. Increasing the cover period also significantly reduces emissions compared to shorter cover times. If a virtually impermeable film is placed over the soil for 30 d, emissions can be nearly eliminated, although this may be impractical in current agricultural systems. B. Soil Conditions 1. Soil Water Content Increasing soil water content has been considered as a means for controlling MeBr movement (Goring 1962; Reible 1995; Jin and Jury 1995). The effect of water content on MeBr volatilization can be explained by the interactions of soil water content and the retardation factor, Rg = (θ + Kdρb)/ Kh + ε, and tortuosity factor, (e.g., τ = (θ − ε)10/3/η2) in MeBr gas-phase transport, where η, θ, ρb, Kd, and Kh, respectively, are the porosity, water content, bulk density, liquid–solid, and liquid–gas partition coefficients, and the air content, ε, is ε = η − θ. When the water content in laboratory columns containing Greenfield sandy loam was increased from 0.058 to 0.180 cm3 cm−3, Rg increased from 1.21 to 1.58, τ decreased from 0.241 to 0.076, and the effective soil diffusion coefficient was reduced by 76%. For volumetric water contents of 0.058 and 0.124 (cm3/ cm3), estimated emission loss after correcting for the presence of the column bottom was approximately 77% of the applied MeBr (Gan et al. 1996; Table 9); this indicates that at moderately dry soil conditions, water content does not strongly affect emissions. When the water content was increased to 0.180 cm3 cm−3, 62% of the applied MeBr was lost (Table 9). As the soil water content increased, the maximum MeBr flux density decreased and the time interval before reaching the maximum flux density increased. Measurements of the MeBr gas concentration in the soil also indicated rapid movement through the soil column for the drier soils. MeBr in these soil columns was completely depleted 54–72 hr after application. For the wetter soil, measurable concentrations remained in the column until 144 hr after application. In a field experiment (Yates et al. 1997), lower MeBr emissions were observed for bare soil and deep application than for a tarped, shallow application in the same field. Part of this difference may be attributed to the higher water content of the soil profile. During the deep-injection study, average soil water
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Table 9. Effects of soil type, water content, and bulk density on methyl bromide dissipation.
Treatment Water content, (L3 L−3) 0.058 0.124 0.180 Bulk density (g cm−3) 1.40 1.70 Soil type Greenfield SL Carsetas LS Linne CL aCorrected
Total emissions (measured) (%)
Total degradation (%)
Mass balance (%)
Total emissions corrected using diffusion modela (%)
90 90 75
6 12 26
96 102 101
77 77 62
90 64
12 29
102 93
77 53
90 90 44
12 9 49
102 99 94
77 77 37
to account for the presence of a lower impermeable boundary, see section VII.C.1.
content around the injection point (68 cm depth) was 0.223 (cm3 cm−3), whereas that observed during the shallow-injection study was 0.145 cm3 cm−3. Further, light rain occurred during the early part of the experiment, which helped to seal the soil pores. Although measured water content data were not given, Yagi et al. (1995) also attributed the decrease in MeBr emission from 87% in their first study to 34% in their second study, in part, to soil moisture differences. Similar results were observed in the laboratory by Jin and Jury (1995). 2. Soil Bulk Density Soil bulk density can also affect MeBr transport because the pore space decreases as bulk density increases. Bulk density, ρb, is related to porosity, η, from the relationship η = (1 − ρb /ρp), where ρp is particle density. In laboratory columns packed with Greenfield sandy loam (Gan et al. 1996), the corrected cumulative volatilization loss for a column with a bulk density of 1.70 g/cm3 was 53%, significantly lower than the 77% loss from a column with a bulk density of 1.40 g/cm3 (Table 9). The columns with higher bulk density behaved in a manner similar to the wetter soil column described earlier. Measurable volatilization continued for 120 hr, the maximum flux density was reduced ⬃60% compared to the low bulk density column, and the time to reach maximum flux increased to 6.5 hr after application, compared to 2.5 hr in the low bulk density column. In the untarped, deep-injection field study (Yates et al. 1997), the field was disced and packed approximately 5 min after MeBr injection. Discing and sur-
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face packing closed the openings above injection fractures and increased bulk density near the surface; this, along with a higher water content, probably contributed to the reduced total emission compared to the shallow-tarped experiment. Packing the soil surface and carefully closing the soil fractures created during application should be considered for minimizing MeBr volatilization. C. Effect of Degradation Rate on Emissions Hydrolysis and methylation are the principal degradation processes removing MeBr from agricultural soils (see Section VI.A). Degradation affects volatilization because it removes MeBr from the soil, making it unavailable for transport to the atmosphere. Total degradation is affected by both intrinsic soil degradation and the type and performance of agricultural films used for containment because this increases soil residence time. Predicted total MeBr emissions as a function of the degradation half-life are shown in Fig. 9 when MeBr is injected at 25 cm and the soil surface is covered with either a HDPE or Hytibar film for 5 or 30 d. For a wide range in soil degradation half-lives, a significant reduction in emission occurs when using a virtually impermeable film (e.g., Hytibar) together with long cover periods. This behavior is also shown in Fig. 8. Such information can aid in reducing emissions by providing a prediction of expected emissions to allow for comparison between various application methods. This information can also be used to allow specification of target total emissions and provide information needed to achieve this goal. For example, the dotted line in Figure 9 gives the soil degradation half-life necessary to achieve 20% emissions using Hytibar film and a 30-d cover period. A soil with a shorter life-life would not exceed the 20% emissions threshold. It is clearly shown in Fig. 9 that a very
Fig. 9. Total emissions of MeBr after injection at 25 cm and covering the soil surface with HDPE or Hytibar film. Simulation parameters: Ds = 1450 cm2 d−1, h = 18, and 0.09 cm d−1 for HDPE and Hytibar, respectively. Dotted line shows the degradation half life needed to achieve 20% total emissions for MeBr application using Hytibar and a 30-d cover period.
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short degradation half-life would be required to reduce MeBr emissions to 20% for a 25-cm injection with a 5-d HDPE cover period. The effect of soil organic matter on MeBr volatilization has been investigated in laboratory soil columns using three soil types. Greenfield sandy loam had relatively low organic matter (0.92%) and clay content (9.5%) and is representative of many soil types in California. Carsetas loamy sand had a very high sand content and very low organic matter (0.22%) and clay content (0.1%). Linne clay loam was relatively rich in organic matter (2.99%) and clay (25.1%). Soil type had a pronounced effect on MeBr volatilization behavior, as shown in Table 9. Volatilization from untarped Carsetas and Greenfield soil columns following 30-cm injection was very rapid; corrected cumulative emissions were 77% of the applied for both soil columns. However, under the same conditions with the Linne clay loam, only 37% was lost. Analysis of Br− concentration in soil at the end of the experiment revealed that 49% of the applied MeBr was degraded to Br− in the Linne soil, whereas the degradation in Carsetas and Greenfield soils accounted for approximately 10% of the applied MeBr mass (Table 9). The enhanced degradation in Linne clay loam is due to its higher organic matter content as indicated by earlier work (Brown and Rolston 1980; Arvieu 1983; Gan et al. 1994). Using a gas-phase diffusion model, Reible (1994) predicted that when the soil organic carbon content was increased from 2% to 4%, the MeBr emission rate decreased from 45% to 37% following a tarped (2-d), 25-cm application under the assumed conditions. However, in his simulation, only the effect of soil organic matter on adsorption behavior was considered. From the column experiments, it is clear that enhanced degradation resulting from higher organic matter may play an important role in reducing MeBr volatilization in organic matter-rich soils. Increasing the rate of MeBr transformation in soil by addition of organic amendments or nucleophilic compounds has also been demonstrated to reduce MeBr emissions (Table 10, Fig. 10), as demonstrated using laboratory soil columns (Gan et al. 1998b,d). Adding organic amendments, such as composted manure, to soil results in more rapid transformation of MeBr because of an increase in both abiotic and biological transformation rates (Gan et al. 1998b). Incorporation of organic amendments into the surface soil resulted in reduced emissions (Gan et al. 1998b). With no amendment, cumulative volatilization Table 10. Cumulative fumigant emissions (% of applied) observed in unamended soil columns and columns that included composted manure (5%) in the surface 5 cm of soil (amended). Compound Methyl bromide 1,3-Dichloropropene Methyl isothiocyanate
Unamended
Amended
68.2 25–34 21.3
56 14–18 0.3
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Fig. 10. Total emissions are reduced after application of ammonium thiosulfate (ATS) to the soil surface.
losses were 68%. Incorporating composted manure (5%) into the top 5 cm of soil resulted in cumulative MeBr emissions of 56% (Table 10). Increasing the proportion of organic amendment at the soil surface further decreased the volatilization loss, and mixing 20% composted manure into the top 10 cm of soil resulted in cumulative emissions of 40%. Application of composted manure also decreased emissions of MeBr alternatives (Table 10). Addition of nucleophilic compounds such as fertilizer ammonium thiosulfate (ATS) greatly enhances the rate of MeBr degradation in soil (Gan et al. 1998d). Addition of such compounds at the soil surface can provide an effective barrier to its volatilization while maintaining adequate concentrations in the root zone to provide efficacy against soil-borne pests. Total emissions were reduced from 61% (unamended column) to less than 10% after adding ATS to the soil surface at a 3 : 1 molar ratio (see Fig. 10). A field study showed that adding ATS (at 640 kg/ha) to the soil surface had no discernable effect on the efficacy of MeBr for controlling nematodes and weeds (Gan et al. 1998d). Emissions of MeBr alternatives were also reduced with application of ATS and organic amendments (Gan et al. 2000a). Wang et al. (2000) found that the byproducts of ATS degradation of fumigant compounds are not toxic. Because thiosulfate salts are inexpensive materials used as fertilizers, this approach has promise for field application.
XI. Considerations for Developing Alternatives to Methyl Bromide To be effective, a soil fumigant should be able to penetrate readily in a soil environment and reach the pests in targeted soil zones. As listed in Table 11, the banned and remaining registered fumigants have markedly higher vapor pressures, lower boiling points, and higher air–water partitioning coefficients (KH) than a typical nonvolatile pesticide (e.g., atrazine). Thus, these compounds have much higher mobility in the vapor phase than nonvolatile pesticides and
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Table 11. Boiling points, vapor pressures, and Henry’s law constants (KH) of selected pesticides.
Compound Methyl bromide Chloropicrin cis-1,3-Dichloropropene trans-1,3-Dichloropropene Methyl isothiocyanate (MITC) Ethylene dibromide (EDB) Dibromochloropropane (DBCP) Atrazine
Boiling point (°C) 3.6 112 106 111 119a 132 199 171–174 (Melting point)a
Vapor pressure at 20 °C (mm Hg)
KH at 20 °C
1380 20 25 18.5 20b 7.7 0.58 2.89 × 10−7 b
0.24 0.093 0.056 0.041 0.010c 0.023 0.006 1.0 × 10−7 c
aMerck
Index (1996). et al. (1992). cEstimated from vapor pressure and solubility data from Wauchope et al. (1992). All other data from Goring (1962). bWauchope
are rapidly dispersed through soil (see Table 4). The high diffusive mobility of a fumigant in soil means that it has a greater tendency to volatilize into the air and/or reach the groundwater than a nonvolatile pesticide, which is evidenced by the wide detections of ethylene dibromide (EDB) and dibromochloropropane (DBCP) in groundwaters, and significant emissions of MeBr and 1,3-dichloropropene into the atmosphere. Although the potential for volatilization or groundwater contamination is largely controlled by the inherent vapor pressure, as discussed extensively here in the case of MeBr, many application, soil, and meteorological factors affect the fundamental processes (e.g., transformation and transport) that dictate pesticide fate and environmental contamination. The few remaining fumigants that are likely to replace MeBr include 1,3dichloropropene (1,3-D), chloropicrin, and methyl isothiocyanate (MITC) precursors (e.g., metam sodium, Dazomet, Basimid). After application to the soil, MITC precursors react with water to form the active ingredient MITC. Unfortunately, all these candidates have narrower spectra of activity than MeBr, and it is likely that combinations of chemicals will have to be used to achieve broad control. For example, Telone-C17, a mixture of 1,3-D containing 17% chloropicrin, was demonstrated to be a promising alternative to MeBr in controlling a wide variety of soil-borne pests (Melichar 1995). Some field monitoring and model simulation studies indicate that significant volatilization losses of 1,3-D and MITC can occur under current practices. Air pollution was the cause of a temporary suspension of Telone-II, the main commercial formulation of 1,3-D, in California between 1991 and 1994. Both 1,3-D and MITC are classified as air toxic compounds in the USEPA Clean Air Act, and their emissions into the air at excessive levels will likely cause concern.
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As in the case of MeBr, most of the application methods used for applying 1,3-D, MITC and chloropicrin were originally developed based on performance or efficacy. The relationships between these application techniques and the environmental consequences (e.g., volatilization and groundwater contamination) have not been fully investigated. Many of the methods demonstrated to reduce MeBr emissions also have the potential to reduce emissions of MeBr alternatives. For example, tarping the soil surface with HDPE and virtually impermeable films reduces emissions of methyl iodide (Gan et al. 1997b) and chloropicrin (Gan et al. 2000b). Surface application of organic amendments at the soil surface reduces emissions of MITC (Gan et al. 1998b), and ATS application reduces emissions of 1,3-D with no change in efficacy (Gan et al. 2000a). To preserve these few remaining fumigants, it is imperative that further studies be conducted to optimize these application methods so that their adverse effects on the environment are minimized.
Summary The great variation among results of recent experiments measuring the total emission of MeBr from fields implies that many factors influence MeBr transport and transformation in the soil–water–air system and its ultimate loss from the soil surface. It has been demonstrated that variables related to application methods (e.g., injection depth, use and type of surface tarp), soil properties (e.g., water content, bulk density, soil organic matter), and climatic conditions (e.g., air temperature, wind speed, barometric pressure) have pronounced effects on MeBr volatilization following soil injection. The following conclusions can be drawn from this experimental information. Tarping consistently increased the residence time and concentration of MeBr residing in the soil. Prolonged retention of MeBr in the soil resulted in more extensive degradation and reduced cumulative emissions. Research indicates that the polyethylene film typically used for the surface cover is relatively permeable to MeBr and allows significant emissions compared to virtually impermeable plastic films. This effect is more pronounced during periods of high temperature. Soil type, soil water content, and bulk density are important factors affecting MeBr transport and transformation in soil, which ultimately affect volatilization. The total volatilization from a soil with high organic matter content may be drastically reduced relative to that from a low organic matter soil. Amendment of the surface soil with organic matter or nucleophilic compounds that promote increased degradation may offer another method for reducing volatilization. MeBr volatilization may also be decreased by increasing soil water content and bulk density, mainly because of the reduced gas-phase diffusion resulting from reduced soil air-filled porosity. To minimize volatilization, MeBr should be applied during periods of cool temperature, injected relatively deep in organic-rich, moist soil, and the soil surface packed and tarped immediately after the application. Depending on sitespecific conditions, a new high-barrier plastic should be used. Injecting MeBr
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during periods of warm temperature, at a shallow depth in dry, loose soil without the use of low-permeability plastic barriers, will likely result in maximum volatilization rates and therefore should be discouraged. Before adopting any new emission reduction technology, the pest control characteristics of the new methodology should be assessed under soil and environmental conditions typical of the region to optimize efficacy while minimizing environmental contamination. There is considerable current scientific evidence indicating that eliminating MeBr use for soil fumigation may not have a significant impact on stratospheric ozone depletion. Management practices can and have been developed that essentially eliminate atmospheric emissions of MeBr and other fumigant compounds following soil application. Some scientists have suggested that there are natural buffers and various unknown sources of MeBr that make it impossible to ascertain that eliminating soil fumigation with MeBr will significantly improve stratospheric ozone levels. It is quite certain, however, that the phase-out will make it much more difficult for growers to economically provide an adequate and healthful food supply in the U.S. and elsewhere in the world. As the phaseout date approaches, there remains a great need for information about MeBr and stratospheric ozone depletion. Stratospheric ozone must be protected, but recent experiments suggest that it can be protected while still allowing MeBr to be used for soil fumigation. A new approach may be warranted in which state and federal regulations recognize that every chemical is a potential environmental contaminant, depending on the properties of the chemical and the environmental conditions prevailing following its application. Ideally, regulations should incorporate incentives to develop technology that minimizes the likelihood that a chemical becomes an environmental and/or public health problem. Rather than instituting an irrevocable ban, allowing for a suspension of chemical use until the appropriate technology is developed to control the undesirable characteristic(s) of the chemical use would provide much more flexibility to growers and may enhance environmental protection by adopting a proactive approach in which growers, chemical manufacturers, regulators, and the public can have confidence.
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Springer-Verlag 2003
Disposal and Degradation of Pesticide Waste Allan S. Felsot, Kenneth D. Racke, Denis J. Hamilton Contents I. Introduction ....................................................................................................... II. Unique Characteristics of Pesticide Waste ...................................................... A. The Dilemma of Pesticide Mixtures ........................................................... B. Pesticide Persistence at High Concentrations ............................................. C. Regulatory Constraints for Waste Disposal ................................................ III. A Strategy for Disposal .................................................................................... IV. Characterization of Waste Constituents ........................................................... V. Disposal of Unused Pesticide Stocks ............................................................... A. FAO Guidelines on Disposal of Bulk Quantities of Pesticides in Developing Countries .......................................................................... B. Disposal by Incineration .............................................................................. C. Destruction of Chlorinated Pesticides by Potassium Polyethylene Glycol Ether ............................................................................................. VI. Options for Disposal of Pesticide Containers and Packaging ......................... A. Recycling Options ........................................................................................ B. Incineration of Plastic and Paper Packaging ............................................... VII. Options for Management and Disposal of Wastewater ................................... A. Treatment and Disposal ............................................................................... B. Physical Treatment ....................................................................................... C. Chemical Treatment ..................................................................................... VIII. Cleanup of Contaminated Soils ........................................................................ A. Soil Separation Techniques ......................................................................... B. Physical and Chemical Treatment ............................................................... C. Biological Treatment .................................................................................... IX. Conclusions ....................................................................................................... X. Consensus Recommendations of the IUPAC Commission on Agrochemicals and the Environment .................................................................................... A. General Waste Management Strategy ......................................................... B. Obsolete Pesticide Stocks ............................................................................ C. Small Waste Generator Practices for Containers and Rinsewater Disposal ....................................................................................................
Communicated by George W. Ware. A.S. Felsot ( ) Washington State University, 2710 University Drive, Richland, WA 99352, USA Kenneth D. Racke Dow Agrosciences, 9410 North Zionsville Road, Indianapolis, IN 46268–1053 USA D. Hamilton Department of Primary Industries, G.P.O. Box 46, Brisbane, Queensland 4001, Australia
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D. Cleanup of Contaminated Soil .................................................................... E. Future Research Needs ................................................................................ Summary .................................................................................................................... Acknowledgments ...................................................................................................... References ..................................................................................................................
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I. Introduction All countries have had to face problems of pesticide waste, although the nature of the problems and therefore the focus of attention are somewhat different between the highly industrialized countries and the developing countries. In the United States and Europe, for example, pesticide waste management has focused on handling of wastewater, disposing and recycling of containers, and remediation of contaminated soils (Bourke et al. 1992; Dyer 1998). For developing countries, disposal of unused, unwanted, or obsolete pesticide stocks has been a major problem (Jensen 1992). This problem has also created contaminated soils. Regardless of a country’s stage of development, literally every aspect of pesticide technology creates waste. Characterization of the activities generating waste and the identities and quantities of common pesticide wastes have been recently reviewed (Felsot 1998a). Whatever the source of the waste, its characteristics are similar—high concentrations of a diversity of pesticide active ingredients dissolved in either formulation solvents or water containing emulsifiers and surfactants. In addition to potentially contaminating soils directly, liquid wastes contain soil and engine oils or fuel residues, potentially complicating disposal and recycling efforts. At manufacturing sites or wood treatment facilities, wastes tend to be comparatively more homogeneous than those generated by activities directly related to farming, but concentrations tend to be extremely high in comparison. Whether arising by leakage from deficient and worn packaging, improper disposal of rinsewater on land, or leaky spray nozzles, high concentrations of pesticide residues can be extraordinarily persistent (Dzantor and Felsot 1991; Gan and Koskinen 1998), raising the probability for contamination of water resources. Thus, speedy and efficient remediation is warranted to protect environmental quality and human health. Earlier literature about pesticide waste disposal options focused on the pesticide manufacturing industry (Atkins 1972). Thirty years ago, air pollution and land disposal regulations were more primitive in comparison to today’s standards, and options such as burial and ocean dumping were feasible. Although incineration was the most widely used disposal option, chemical and biological degradation processes were used to a limited extent. Variations of these latter disposal techniques have been studied more recently from a basic perspective and could be amenable to small-scale waste generated by individual farmers or commercial agrochemical applicators. Comprehensive reviews of a wide variety of remediation techniques specifi-
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cally for pesticide waste have been published (Krueger and Seiber 1984; Bridges and Dempsey 1988; Enlow 1990; Norwood 1990; Norwood and Randolph 1990; Bourke et al. 1992; Kearney and Roberts 1998). During the 1990s, the Commission on Agrochemicals and the Environment1 within the Division of Chemistry and the Environment of the International Union of Pure and Applied Chemistry (IUPAC) began a critical analysis of strategies and methods for disposal of pesticide-contaminated wastes. Because the Commission has been committed to providing information about agrochemical technology and the environment to developing countries, one objective of the analysis was to elucidate the applicability of waste disposal technologies for small waste generators. Thus, this review extends and updates other reviews (through 2002) but also focuses on small-scale generators of waste and on techniques that are comparatively inexpensive and employ materials that should be readily available. This analysis of pesticide waste disposal and degradation starts with an overview of the characteristics of pesticide waste, including governmental regulations, that raise constraints for disposal. Next, we review practical options for disposing of unwanted pesticide stocks, handling of spray tank rinsewater, and cleanup of contaminated soils. Many techniques are still under development but some seem to have potential for quick implementation on individual farms or perhaps by farm cooperatives. For the various techniques, some assessment of feasibility for small waste generators is given along with the current status of commercialization. Waste disposal practices of chemical manufacturers and product formulators are normally under direct governmental regulations, but these industries should be able to afford sophisticated technologies. Finally, the review concludes with the Commission’s consensus recommendations for a general waste management strategy, handling of obsolete stocks, practices for containers and rinsewater disposal, cleanup of contaminated soils, and future research needs.2
II. Unique Characteristics of Pesticide Waste Almost all studies of pesticide environmental behavior and fate have focused on residues resulting from agronomic rates of application. Residues of pesticides in waste-contaminated soils and water, or alternatively as unused formulated product, differ from normal application residues by their presence at significantly higher concentrations. Compounding the problem of waste disposal and remediation is the routine occurrence of a diversity of chemistries. Furthermore, governmental regulations may restrict how a waste is treated and its final disposition. Any proposed technique for remediation must be designed and imple-
1
The Commission was renamed the Advisory Committee on Crop Protection Chemistry in 2002. Chemical names and CAS numbers of all pesticides mentioned in this review are shown in Table 13.
2
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mented with recognition of the unique characteristics of waste with respect to persistence, chemical diversity, and regulatory restrictions. A. The Dilemma of Pesticide Mixtures Studies of pesticide behavior in water and soil almost invariably focus on one compound at a time. Soil and water affected by pesticide waste may contain a diversity of product chemistries that reduces the efficiency of any one disposal and degradation option. The magnitude of the problem is apparent from surveys of pesticide types and concentrations at a diversity of contaminated sites (Table 1). The complexity of the mixtures of chemicals also demands an accurate qualitative as well as quantitative characterization of the waste or contaminated site before selecting a disposal option. Obsolete stocks notwithstanding, many problems typically start with equipment rinsewater, although spills will certainly occur during spray tank filling. Depending on the crop, one or more pesticides of different types are involved.
Table 1. Number of pesticides analyzed and maximum concentrations reported in different waste forms at selected user sites. Number Maximum of concenpesticides tration
User site (n = total sites studied)
Form of waste
Commercial agrichemical application facility (n = 4)
Equipment rinsewater
7
285
Pesticide products burial site (n = 2)
Soil
9
30
Applicator waste disSoil posal evaporation bed (n = 1)
20
2245
Winterlin et al. 1989
2
106
Bicki and Felsot 1994
Reference Taylor et al. 1988
Hourdakis et al. 2000
Pesticide storage warehouse fire (n = 1)
Soil
Commercial agrichemical application facility (n = 20)
Soil, gravel, water
24
75900
Habecker 1989
Commercial agrichemical application facility (n = 18)
Soil, gravel
3
11757
Andrews Environmental Engineering 1994
Commercial agrichemical application facility (n = 56)
Groundwater
16
Aerial applicator (n = 2) Equipment rinsewater
6
12
1195
Long 1989
Woodrow et al. 1989
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Indeed, for field crops, use of herbicide mixtures has become routine. Improper disposal of rinsewater quickly becomes a problem of soil, gravel, and perhaps water contamination. At this point, remediation becomes more difficult because of the mixture and complexity of the contaminated matrices. As is discussed here, incineration is probably the one technique that would be applicable to a wide range of pesticides and matrices, but this disposal method is not a feasible option for small farmers because of its expense. Thus, development of techniques applicable to small waste generators needs to be tested using a wide variety of pesticide classes. B. Pesticide Persistence at High Concentrations Pesticide dissipation from soil has generally been observed to slow significantly as concentration increases above a certain level (Gan and Koskinen 1998). The pesticides for which this change in persistence has been observed tend to be subject to biodegradation at agronomic rates of application (Table 2). The reason for the increase in estimated time to 50% dissipation at high concentrations is unknown, but at least one hypothesis has suggested microbial toxicity may have inhibited degradation (Dzantor and Felsot 1991; Gan et al. 1995). Prolonged persistence of pesticides increases the opportunity for runoff and leaching. In-
Table 2. Effect of concentration in soil on time to 50% dissipation (DT50%) of selected pesticides. Concentration (mg/kg)
DT50%(wk)
Location
10 10,000 4 4,300
<2 >40 <2 >52
Lab
Gan et al. 1995
Field
Gan et al. 1995
7 6,400 5 5,000
<4 >24 <5 <12
Field Field Lab Lab
Gan et al. 1996
10 10,000
<4 >52
Lab Lab
Dzantor and Felsot 1991
Trifluralin
100 1,000
116–189 168–544
Lab Lab
Schoen and Winterlin 1987
Captan
100 1,000
1–2 30–48
Lab Lab
Schoen and Winterlin 1987
Diazinon
100 1,000
23–30 77–160
Lab Lab
Schoen and Winterlin 1987
Pesticide Alachlor
Atrazine
Metolachlor
Reference
Gan et al. 1996
Concentrations 10 ppm or less are typical of broadcast applications incorporated into the upper 15 cm or less of soil.
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deed, leaching potential seems to be greater with higher concentrations than comparatively lower concentrations (Davidson et al. 1980; Gan et al. 1995, 1996; Gan and Koskinen 1998). C. Regulatory Constraints for Waste Disposal The plethora of worldwide regulations makes it difficult to generalize about regulatory constraints for disposal of pesticide waste. Federal or country-wide regulations may uniformly dictate how waste is handled, but state or regional laws may add idiosyncrasies that further restrict pesticide manufacturers and users. One dilemma facing many countries is making a clear distinction between pesticide waste and dangerous or hazardous waste. A survey of pesticide waste regulations across the world is beyond the scope of this review, but two case studies in the U.S. and Australia serve to illustrate some of the issues facing pesticide waste generators. 1. U.S. Regulations. In the U.S., pesticide waste comes under the jurisdication of three laws, FIFRA (Federal Insecticide, Fungicide and Rodenticide Act), RCRA (Resource Conservation and Recovery Act), and CERCLA (Comprehensive Environmental Response, Compensation and Liability Act) (Fitz and Jensen 1998). All pesticides are registered by the U.S. Environmental Protection Agency (EPA) under authority of FIFRA, and the mechanism for controlling pesticide use is the language of the product label. For example, the label includes information about crops or sites to which the pesticide can be applied, how the pesticide can be applied, the rate and frequency of application, use restrictions (e.g., do not apply near water, avoid drift), warnings about health and environmental hazards, and pesticide and container disposal. FIFRA mandates manufacturers to provide EPA with information for safe disposal, but as of 1998, EPA had not routinely required this information (Fitz and Jensen 1998). Although product labels require information about the transportation, storage, and disposal of the pesticide product and its container, the content is superficial and the language may be ambiguous or confusing (Lounsbury 1992). Furthermore, in addition to not providing enough information, the label may have statements that are in contradiction to state laws. Although EPA is mandated to prescribe procedures and standards for the removal of pesticides from containers before disposal, FIFRA also states that the regulations shall not diminish the authority of the Solid Waste Disposal Act that was amended to become RCRA. One resulting conflict has been deciding when a pesticide is no longer regulated as a product under FIFRA but as a waste under RCRA. EPA has interpreted the law to mean that a pesticide stops being a product when a decision is made to discard it. Applicability of specific RCRA provisions to pesticide waste disposal is determined by whether the discarded pesticide is considered a solid waste and meets the test for being classified as a hazardous waste. Solid wastes are products that are discarded, defined as abandoned (disposed of, burned, incinerated,
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or accumulated, stored, or treated before abandonment), recycled (in a manner different from the product’s registration), or inherently “waste-like.” Recycling includes application to land in a manner constituting disposal, but pesticides are exempt if applied to the land consistent with their ordinary manner of use. Discarded pesticides that are solid wastes are also hazardous wastes if they meet the characteristics of ignitability, corrosivity, reactivity, or toxicity, or they are specifically listed as hazardous. Listed hazardous wastes include banned active ingredients such as 2,4,5-T and pentachlorophenol as well as many currently registered products when they are discarded. If soil becomes contaminated with hazardous waste, then the medium itself is considered hazardous. Contaminated media must be handled as hazardous waste until which time they no longer contain the waste. The EPA and authorized states have the discretion under RCRA to determine on a case-by-case basis or under some guidance criteria when the concentration of the hazardous waste is so low as to be interpreted as being absent from the medium. Generators of hazardous waste are subject to a set of “cradle-to-grave” regulations governing how the waste is handled and documentation of every facet of storage, treatment, and transportation. A specific exemption is provided for farmers who triple rinse all containers and dispose of the pesticide residues on their own land in a manner consistent with the disposal instructions of the product label. A related but confusing issue involves remediation of contaminated media. For example, product may spill or be found at a mixing and loading site after years of use. The soil could feasibly be excavated and applied to farmland in a manner consistent with the permitted crop and at the permitted application rate, but whether the pesticide is a waste or a product is still open to EPA and state interpretation (Fitz and Jensen 1998). Although the rules seem somewhat confused for deciding when a pesticide product becomes a waste and when contaminated media should be treated as hazardous, obsolete stocks are considered hazardous waste if they meet the definitions of ignitability, corrosivity, reactivity, or toxicity. Promulgation of the Universal Waste Rule (UWR) relaxed the stringent requirements for small generators of hazardous wastes that fell into the category pesticide, battery, or thermostat (Centner 1998). The UWR allowed States to hold amnesty collection days in which farmers could bring obsolete stocks to a central location for packaging and proper disposal without being burdened by the usual record keeping required of larger hazardous waste generators. The receivers of the universal waste can accumulate no more than 5000 kg of waste at one time, but they are responsible for ensuring the waste is transported to a treatment facility. 2. Australian Regulations. Pesticide regulation in Australia is dispersed between the Commonwealth and the States (ARMCANZ 1998). Recently (1995), a National Registration Scheme for Agricultural and Veterinary Chemicals became operational. The National Registration Authority (NRA) has the responsibility to assess the health and environmental issues before registering a pesticide. On approval of a registration, the NRA issues a product label that provides
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users with directions for best agricultural practices to minimize impacts on health and the environment. The NRA essentially controls pesticides at the point of manufacture, distribution, and supply, including retail sale. The Australian States and Territories are responsible for controlling the use of pesticides once they have been sold and for guiding users on the use of the registered products. Thus, in Australia, pesticide waste management is a State responsibility, although a national strategy for disposing of used containers has been proposed (ARMCANZ 1998). Various agricultural groups have Codes of Practice that also include pesticide product management and disposal. For example, generic practices include storing pesticides in a safe manner and segregation of products from fertilizers and seeds. Organochlorine and arsenic-containing products must be stored separately with bunding (containment walls). Commodity- or industry-specific Codes of Practice are approved by the State. For example, the fruit and vegetable industry of Queensland has developed a Code of Practice called Farmcare that is sanctioned under the Queensland Environmental Protection Act (EP Act). Compliance is voluntary, but growers have a strong incentive to follow the Code for accreditation purposes and meet their “general environmental duty” as defined under the EP Act. Pesticide waste in soil is regulated under Queensland’s Contaminated Land Act. Land owners or occupants are required to report to the Queensland Department of Environment (QDE) if there is a contaminated site on their land. The QDE decides whether to list the land on the Environmental Management Register (EMR; low probability of human or environmental harm) or the Contaminated Land Register (CLR; remediation required) (Queensland Department of Environment 1998). If no offsite impacts are expected, the site is listed under the EMR and therefore would not require remediation as would a CLR listing. For example, cattle dipping sites would likely be regulated under an EMR because of the low risk of offsite hazards. Wastewater may be allowed to be discharged into sewers during “peak flow” to allow dilution, but such practice is unlikely to be compatible with the Farmcare Code of Practice for appropriate waste management. The primary recommended disposal option is recycling of rinsewater by spraying it onto the field margins. However, Farmcare also recommends, as a last resort, the discharging of unwanted dilute agricultural chemicals into an on-farm disposal pit if the pit meets certain best management practices. For example, the pit should be no more than 1 m deep and at least 2 m above groundwater. The pit should be lined with impermeable soils and located far enough away from dwellings, surface water, crops, and livestock to minimize risk. Disposal of livestock dip or postharvest fruit dip waste has specific guidelines imposed by the Queensland Department of Primary Industries. For example, the guidelines specifically prescribe using slaked or quicklime for disposal of the organophosphorus insecticides dimethoate and fenthion. The Australian regulations seem similar to U.S. regulations with regard to the designation of Scheduled chemical wastes as listed in the Environmentally
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Hazardous Chemical Act of 1985. Scheduled wastes, which include organochlorine insecticides such as chlordane, dieldrin, and DDT, and the phenoxyacetic acid 2,4,5-T must be disposed of by incineration at an approved facility. All the organophosphorus pesticides and the organochlorine endosulfan are not scheduled but require alkaline hydrolysis before treatment at a centralized facility. Given the variation in regulatory approaches between two industrialized countries, a pesticide user should check country, state, and local regulations before proceeding with a pesticide waste management plan.
III. A Strategy for Disposal Obviously, the first priority for any waste management strategy should be practices that prevent or at least minimize generation of waste and promote recycling of materials. A good example of such a strategy is the Australian Waste Management Hierarchy classification (Table 3). Other useful strategies for waste management as well as pesticide stock management have been published by FAO (1996a,b) and GIFAP (1991). The International Group of National Associations of Manufacturers of Agrochemical Products (GIFAP 1991) has published a document for guidance on the selection of practical options for disposal of unwanted pesticide stocks. The document, however, provides a practical strategy for determining how any type of pesticide waste should be managed and remediated. GIFAP’s strategic plan for choosing the most feasible option of remediating waste is summarized in Fig. 1. Once the product is identified, a determination can be made to either use or recycle it. Even wastewater streams can be recycled if the pesticides intended for a specific crop and pest are segregated from one another and adequate storage is available (Taylor et al. 1988). If use and recycling are not feasible options, then treatment or destruction technologies must be employed. At this point, the strategic analysis (“option filter” in Fig. 1) examines the local availability of disposal technologies, technical feasibility, infrastructural and material needs, logistical considerations, monitoring requirements, and longterm needs for disposal. Potential options are then subjected to health and environmental safety assessment (i.e., risk assessment) and a consideration of risk management factors (social, political, and legal constraints). The preferred disposal option is chosen after also factoring in cost. Although small-scale generators of waste cannot be expected to conduct detailed strategic analyses and planning as recommended by GIFAP (1991), agricultural or environmental regulatory agencies can conduct the assessments and then present an array of feasible choices to the waste generator. Smaller generators could also band together in cooperatives to share the cost of disposal. In addition to the GIFAP strategy, pesticide users can follow the Australian Waste Management Hierarchy (see Table 3). The key factor in the GIFAP strategy and Waste Management Hierarchy is that disposal is decentralized, that is, it is geared to local conditions and not dictated by high technology that only industry in the most developed countries can afford.
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Table 3. Waste management hierarchy promulgated by the Queensland Farmers’ Federation Environmental Code of Practices for Agriculture and gazetted as an approved code of practice under Section 219(1) of the Australian Environment Protection Act 1994 (Queensland Farmers’ Federation 1998). Classification
Practice
Objective
Least preferred
Waste disposal
The last waste management option that should only be adopted when all other practical possibilities have been considered. Waste should be confined to a defined area and managed to minimize harm to the environment or public health.
Good
Waste treatment
Using a process that turns waste into a form more easily used or more easily or safely disposed of (e.g., adsorption by activated carbon, chemical or biological detoxification).
Better
Waste recycling
Strategies to reuse, reprocess, or recover a product (e.g., using returnable chemical containers; accessing programs that use crushed chemical drums for industrial fuel; spraying rinsewater on crop edges).
Even better
Waste reduction
Reducing the amount of waste created by using whatever means are available (i.e., changing practices; e.g., ultralowvolume applications; use of low-rate compounds).
Most preferred
Waste avoidance
Avoiding generation of waste at the source by looking at the situation and changing practices or by choosing the least hazardous products.
IV. Characterization of Waste Constituents The specific constituents in waste can dictate disposal options. For example, old pesticide stocks containing lead arsenate or mercury are not good candidates for incineration. Inert ingredients such as solvents and surfactants may affect the potential for bioremediation. Thus, qualitative as well as quantitative analyses are the first steps before planning a disposal option. Although industrialized countries are likely to have readily accessible equipment such as gas and high pressure liquid chromatographs outfitted with either selective detectors or coupled with a mass spectrometer (HPLC, GC/MS), developing countries may have trouble finding competent facilities nearby that are also affordable. Affordability
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Fig. 1. Strategic plan for selecting an option for pesticide waste disposal. (From GIFAP 1991.)
is a key impediment for any kind of small waste generator to obtaining proper waste characterization. Appropriate technology for pesticide analysis in developing countries will depend on the level of government commitment to long-term funding and the availability of well-trained dedicated analysts (Akerblom and Cox 1996). Qualitative and quantitative analysis of pesticides in obsolete stocks or contaminated soils can be achieved without resorting to dedicated HPLC and GC/MS. Thinlayer chromatography (TLC), deployment of immunoassays (e.g., enzyme detection using acetylcholinesterase assays), and colorimetric assays are well suited to the high concentrations associated with waste. Using a combination of techniques, identification of inorganic and organic analytes can be confirmed. TLC may be the most amenable technique for laboratories in developing countries. TLC plates can be made onsite, avoiding the more expensive option of purchasing precoated plates. One study has examined the Rf of more than 188 pesticides in up to eight single solvent systems (Ambrus et al. 1981). The study also assessed six detection reagents, including one bioassay with fungal spores and a biochemical assay with cholinesterase enzyme inhibition. The most versatile developing solvents were ethyl acetate and methylene chloride. Each detection reagent had sensitivity for one or more groups of chemically analogous
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pesticides. Standard compounds could be cochromatographed with the unknowns to standardize each TLC run, and known analytes could be simultaneously run for identification by Rf matching. Purchase of purified organic solvents might be cost prohibitive for some laboratories, but the reagent grade solvents could be distilled to a higher purity (Akerblom and Cox 1996).
V. Disposal of Unused Pesticide Stocks The Food and Agriculture Organization (FAO) (1996a) has defined obsolete pesticide stocks as products that can no longer be used for their intended purpose or any other purpose and therefore require disposal. Products become obsolete as a result of registration cancellations or other prohibitions, deterioration as a result of improper or prolonged storage, and lack of other feasible uses. Product deterioration is indicated by chemical and/or physical changes that may cause an unacceptable hazard to crops, the environment, or human health. The product may have undergone an unacceptable loss of biological activity. The physical properties of the formulation may have changed to the point of making application infeasible. Unused or obsolete pesticide stocks are a major waste problem in developing countries that ironically originated with the good intentions of developed countries to donate pesticides. Although the U.S. has banned the export of obsolete pesticides to developing countries (Jain 1992), excessive past donations from developed countries have resulted in thousands of metric tons of obsolete stocks scattered among numerous countries, especially in Africa (Jensen 1992). In 1996, an inventory of obsolete pesticides among 28 African countries had estimated about 15,000 tonnes needed management (FAO 1996a). The total of obsolete stocks in non-OECD (Organization for Cooperation and Economic Development) countries has been estimated to be in excess of 100,000 tonnes. Obsolete stocks often possess badly deteriorated labels and are inadequately stored in corroded and leaking containers, resulting in contaminated soils as well as adulterated product. Even in developed countries, obsolete pesticides are problematic, but as mentioned under the discussion of regulations, the U.S. has relaxed hazardous waste rules to deal with these cases. In the U.S., collections of unused pesticides are sponsored by various agencies of state government under the aegis of the Universal Waste Rule (Lounsbury 1992; Centner 1998). The rule recognizes the problem of the prevalence of many persons with relatively small quantities of hazardous waste (Centner 1998). The efficacy of these programs depends on a combination of several factors—whether the program is specifically agricultural in scope, has permanent funding, and has a high ratio of waste collection to expenditures (Centner 1998). One ongoing problem with state programs is communicating the existence of the programs to farmers, especially in light of the liabilities associated with waste that could be classified as hazardous and thus subject to stringent handling and disposal rules.
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A. FAO Guidelines on Disposal of Bulk Quantities of Pesticides in Developing Countries The FAO has developed guidance documents to manage and dispose of obsolete stocks (FAO 1996a,b). The guidance recommendations can be distilled to four themes: inventory of stocks, site stabilization and temporary storage, disposal, and prevention of accumulation of obsolete stocks. The first step is to determine if the stocks are obsolete through an inventory of stored, unused materials. Labels may be nonexistent or unreadable so that analysis of the contents may be required before a mitigation decision can be reached. Most of the obsolete pesticides in Africa were either banned insecticides such as DDT, dieldrin, and lindane, or organophosphates. Typical analysis would best be handled with GC and/or HPLC, but this instrumentation may not be accessible in many countries. However, owing to the high concentrations of pesticide likely present, simpler techniques employing thin-layer chromatography with UV detection may be suitable, assuming analytical standards for the suspected pesticides are available. Once a determination of the extent and nature of obsolete stocks has occurred, site stabilization to improve the safety of temporary storage is necessary. The FAO (1996b) document provides step-by-step guidance to rectify the situation. Few permanent secure facilities have been constructed as an interim storage solution. A recent report from Poland presented a design for secure storage of unused stocks (Stobiecki et al. 1998). Funded by the European Union, construction of such facilities followed extensive site assessments of old storage areas and then repacking of the contents into sealed drums. Once a storage site is secured, disposal options can be considered. Disposal methods that the FAO considers acceptable and unacceptable are listed in Table 4. Of the methods listed, high-temperature incineration would be the most efficient, but lack of accessibility to the right type of incinerator and cost often make this option impractical.
Table 4. Acceptable and unacceptable methods for disposal of bulk quantities of obsolete pesticides in developing countries. Acceptable methods
Unacceptable methods
High-temperature incineration Chemical treatment Engineered landfill Long-term controlled storage
Open burning Burying or landfill Discharge to sewer Solar evaporation Landfarming Deep well injection Methods developed for soil and groundwater remediation
Source: FAO (1996a).
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United Nations programs have sponsored the packing and shipping of obsolete stocks to western Europe for final disposal (Jensen 1992), and transfer of the waste to an industrialized country may actually be cheaper than disposing of it in country (FAO 1996a). The disposal programs are temporary and dependent on renewed allocations of funding. Thus far, incineration has been the preferred method, perhaps because it is the most effective method to date, especially for persistent, highly chlorinated organics such as DDT and dieldrin. Chemical methods have received attention and hold promise for lesser-developed countries where incineration may not presently be feasible. Incineration and a promising chemical destruction method for chlorinated organics are discussed next. B. Disposal by Incineration 1. Efficiency of Incineration. The option most frequently used for final disposal of unused stocks is generally the same in any country—destruction by incineration. Simply described, incineration is a high-temperature thermal oxidation process whereby a chemical is decomposed into gases and unburned solids known as ash and slag (FAO 1996a). Waste gases containing water, CO2, acids, particles, and metal oxides are emitted through a vent stack. The emissions, however, can be cleaned to various degrees by equipping the incinerator with a scrubber and electrostatic filters. Hazardous waste incinerators achieve high levels of destruction by employing a main chamber and an afterburner. Residence time of the combustion gases ideally should be at least 2 sec with temperatures maintained at a minimum of 1100 °C. Research specifically focused on the chemistry of pesticide waste incineration has been neglected over the last few decades. An early study reported the temperatures of combustion for a wide diversity of pesticides (Kennedy et al. 1969). Temperatures of complete combustion as measured by differential thermal analysis for a number of analytical grade reference standards of active ingredients and their formulations ranged from 545 to 879 °C (Table 5). Differences, although not very large, were noted between the reference compounds and the formulations, suggesting that considerations about combustion efficiency must be given to the inert ingredients in the formulation. In general, there was little change in combustion efficiency between 600 and 1000 °C for a number of pesticides likely to be found in obsolete stocks (Kennedy et al. 1969) (Table 6). Surprisingly, more than 99.5% of recalcitrant organochlorine hydrocarbons such as DDT and dieldrin were combusted at temperatures of 800 °C, but less than 90% of biodegradable compounds such as atrazine and carbaryl were destroyed at temperatures of 1000 °C. In contrast, malathion and fenitrothion were completely destroyed at incinerator temperatures of 730–760 °C, but efficiency was also dependent on transit (i.e., residence) time within the furnace (3–4 sec was optimal) (Ahling and Wiberger 1979). EPA studies showed 99% destruction of 16 currently used insecticides and herbicides at temperatures ranging from 200 to 700 °C (Steverson 1998). Calculated Arrhenius rate constants for the incineration reactions ranged from 7 × 108 to 2 × 1016 sec−1.
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Table 5. Temperatures of complete combustion for analytical reference standards and active ingredient formulations of selected pesticide classes. Complete combustion temperature (°C) Pesticide
Reference standard
Formulation
717 602 650 724 800 560 840 620 775 665 663 613 550 545 879 840
731 623 600 678 596 850 850 640 550 612 715 592 640 646 842 690
2,4,5-T 2,4-D Atrazine Carbaryl DBCP DDT Dicamba Dieldrin Diuron Dimethyl sodium arsonate Malathion Paraquat Picloram Phenyl mercury acetate Trifluralin Zineb Source: Kennedy et al. (1969).
Table 6. Influence of muffle furnace temperature on percentage loss of selected pesticides in commercial formulations. Furnace temperature (°C) Active ingredient
600
700
800
900
1000
Picloram Atrazine DBCP Trifluralin Malathion 2,4,5-T Paraquat Dieldrin DDT 2,4-D Carbaryl
90.8 87.8 99.6 99.7 95.3 99.9 98.3 99.1 99.2 99.8 88.7
91.8 88.1 99.6 99.8 96.0 99.9 98.6 99.4 99.3 99.9 88.8
95.6 88.8 99.6 99.8 96.3 99.9 99 99.5 99.7 99.9 88.8
98.7 88.9 99.6 99.8 96.4 99.9 100 99.5 99.9 99.9 89.1
99.2 89.0 99.6 99.8 96.7 99.9 100 99.5 100 99.9 89.5
Source: Kennedy et al. (1969).
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In developed countries, dedicated hazardous waste incinerators operate under government-prescribed conditions of combustion chamber temperature, residence time, air level, and emission standards. During 1974, the EPA recommended optimal incinerator operating conditions for acceptable destruction of organic pesticides, namely 1000 °C with a 2-sec retention time (Ferguson and Wilkinson 1984). Meeting the specifications of combustion temperature and residence time are important for reducing the production of products of incomplete combustion (PICs). The EPA and private companies have successfully conducted test burns and actual disposal operations (>99.99% destruction and removal efficiency) on a wide range of pesticides (Ferguson and Wilkinson 1984; Oberacker 1988, 1989). The FAO (1996a) has reviewed types of incinerators that might be considered by developing countries for disposing of obsolete stocks. The advantages, disadvantages, and pesticide limitations of the various incinerator options are summarized in Table 7. In developed countries, the most commonly used large-scale incinerator is the rotary kiln (Steverson 1998). Rotary kiln incinerators are quite versatile, with the ability to handle solid and liquid wastes. Incineration is the most efficacious option for disposal but also the most expensive; furthermore, its accessibility to small producers or rural communities may be limited. Large-scale incinerators such as the rotary kiln offer the best option for destruction of waste, but they are likely nonexistent in developing countries. Thus, the FAO recommends that obsolete stocks be packaged and shipped to industrialized countries. Of course, such a process is highly regulated and proper permits must be obtained. Nevertheless, the multicountry and multiagency Pesticide Disposal Project has aided several African and Near East countries for disposal in European hazardous waste incinerators (Guenther et al. 1998). For example, 220 tonnes of dieldrin product and contaminated containers in Mauritania were packed and shipped to Netherlands for high-temperature incineration. This cooperative project between the GTZ (Deutsche Gesellschaft fur Technische Zusammeranarbeit), Shell Chemical Company, Netherlands, and Mauritania was completed at a total cost of $2500–4000 (USD) per tonne. In principle, cement kiln incinerators may be practical in developing countries because they are available, but they would need to be adapted for destruction of chemical waste. Unwanted pesticides can be used as a co-fuel. Temperatures are very high (1400–2000 °C) with long residence times (6–10 sec), and acidic gases and chlorine emissions are adsorbed in cement, obviating the need for special scrubbing systems. Most importantly, considering social and economic constraints, cement product quality is generally not affected (GIFAP 1991; FAO 1996a). Injection systems would need to be modified to accept liquid waste or powdered material. 2. Products of Incomplete Combustion. Incinerator emissions can contain toxic gases and organic compounds, so rigorous use of scrubbers and filters must be used in addition to ensuring that burn conditions are carefully controlled. Some of the emitted by-products include gases such as NOx or bromine, polychlori-
Table 7. Advantages and limitations of incinerator types for destruction of obsolete stocks. Incinerator type Large-scale, fixed
Small-scale, fixed
Cement kiln
Disadvantages
Large capacity (0.5–7 t/hr) with 24 hr op- Expensive (US$10–200 million); requires intensive maneration; maintains temperature of agement and highly trained personnel; must have contin1100–1300 °C; DRE up to 99.99995% uous and substantial waste stream for cost-effectiveness; need landfill for ash and slag residuals Can be installed at the location where the Small capacity (1–2 t/d); expensive (US$1 million); temwaste is generated; may be suitable for perature < 1100 °C; some models have no scrubber; simlocal formulation plants that continuple scrubbers not sufficient for emissions control; ashes ously generate a waste stream must be cleaned out often Requires intensive management and highly trained personnel; must have continuous and substantial waste stream for cost-effectiveness; need landfill for residuals Onsite cleanup; Cost (US$1.5–15 million for unit + $ 1 million for mobiliCapacity to 20 t/d; DRE of 99.999%; zation); requires good road infrastructure; requires elecscrubber system adequate for controltrical power, large quantities of freshwater, and highly ling emissions; can handle liquids, soltrained technicians; landfilling of residuals ids, including soil DRE >99.99995%; Powdered formulations are difficult to inject into the sys1400–2000 °C; residence time in gas tem; cement kilns in developing countries generally not phase 6–10 sec; no scrubber required; suitable; burners must be adopted for liquid wastes no ash disposal; cement quality not af(US$1 million), but powder introduction system of modfected est cost (US$150,000); cannot handle soils; liquids with solids can cause problems; system glitches may cause toxic emissions
Inorganic pesticides, especially those containing mercury If no scrubber, then no chlorine-, phosphorus-, sulfur-, or nitrogencontaining pesticides; If fitted with scrubber, then no organochlorines Organochlorines limited by concentration
Organochlorines limited by concentration; powder formulations
139
DRE: Destruction and removal efficiency. Source: Based on FAO (1996a).
Pesticides not recommended
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nated dibenzodioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs), and polyaromatic hydrocarbons (PAHs). In the U.S., about 3.7 million pounds of the fumigant EDB needed disposal after the EPA cancelled its registration in 1983. Initially, incineration was not considered because of the potential for significant emissions of bromine gas. However, in test burns, an EPA contractor injected 10% dilute sulfuric acid solution into the burner to form HBr, which was then scrubbed out (Oberacker 1989). After cancellation of dinoseb in 1986, EPA studied incineration options for reduction of NOx. NOx was reduced in one test from 2998 ppm after burning in a conventional incinerator to 88 ppm after burning with air staging (diversion of a portion of the air downstream of the flame) and reburning. Air staging (also called staged combustion) converts NOx to N2 (DeMarini et al. 1991). A scrubber removed 97% of the particulate emissions (Oberacker 1989). Another concern about dinoseb incineration was whether nitroaromatic PICs might be produced. Trapping of emissions and measurement of their mutagenicity has been a technique employed to characterize the biological significance of emissions control. In four controlled dinoseb burns, emissions contained mutagenic substances that likely resulted from nitroaromatic compounds (DeMarini et al. 1991). Reburning increased the mutagenic potency by 50%, suggesting that PAHs may have been formed. Mutagenic emission factors derived from dinoseb combustion were several orders of magnitude greater than emissions from industrial utility boilers and power plants but nearly two orders of magnitude less than emissions from wood burning in a controlled-air woodstove. Because many obsolete pesticide stocks are highly chlorinated, PCDDs and PCDFs could be produced; however, these are more likely to be produced at temperatures below 750 °C (Long and Hanson 1983) and in postcombustion chambers with temperatures below 500 °C and abundant particle surfaces (Ghorishi and Altwicker 1995). Currently, there are two hypothesis for formation of PCDDs and PCDFs: de novo synthesis in the presence of HCl and carbon particles or a metal catalyst such as copper, and condensation of chlorophenols (Milligan and Altwicker 1993; Lujik et al. 1994). Combustion of highly chlorinated organic chemicals results in an acidified flue gas as well as chlorinated phenolics, both necessary reagents regardless of the operational mechanism of PCDD/PCDF formation. Scrubbing the flue gas of HCl helps reduce dioxins formation, but the use of carbon filters behind the electrostatic precipitators can lead to their formation, although emissions will be reduced (Lujik et al. 1994). C. Destruction of Chlorinated Pesticides by Potassium Polyethylene Glycol Ether (KPEG) Although incineration is highly efficient for destroying obsolete pesticide formulations, it may not be a feasible option in many developing countries. Many of the obsolete pesticides will likely be either organophosphates and carbamates or chlorinated hydrocarbons and cyclodienes. The former groups can be chemically
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hydrolyzed by raising the pH above 8, but the latter groups are quite recalcitrant to chemical degradation. If the chlorinated pesticides could be dechlorinated first, then they might be amenable to biodegradation, and thus candidates for a land treatment process. During the 1980s, potassium polyethylene glycol ether (KPEG) was found to be capable of dechlorinating PCBs either in solvents or in soil (Brunelle and Singleton 1985; Kornel and Rogers 1985). KPEG reactions also seemed effective for degradation of dioxins and dibenzofurans, known contaminants in older phenoxy herbicide formulations (Teirnan et al. 1989). The reaction is a combination of a phase transfer and nucleophilic substitution at a carbon–halogen bond proceeding by the following mechanism: PEG + KOH → KPEG + H2O
(1)
KPEG + ArCln → ArCln−1 − PEG + KCl
(2)
ArCln−1 − PEG + heat → ArCln−1 − OH + CH2=CH − PEG
(3)
where PEG is polyethylene glycol monomethyl ether and Ar is an aromatic nucleus. The materials needed for the KPEG process should be easy to acquire, and the chemicals are not so exotic as to be unavailable to small-scale waste generators. The reaction vessel can consist of a 55-gal drum with heat tape around it (capable of maintaining at least 70–85 °C) and some kind of electric motor fitted with a mixer (Taylor et al. 1990). With such a vessel and KOH and PEG as the only reagents, at least 98% degradation of phenoxy herbicide waste contaminated with dioxin was achieved. Dioxins were not detectable after a second treatment of the waste. One advantage of using the drum destruction technique is that wastes already packaged in 55-gal drums need not be transferred into another container. If emissions need to be controlled, a vent system with scrubber and absorbent can be easily constructed (Fig. 2). KPEG has also been used to degrade PCB transformer oil, but it has probably been most studied for decontamination of chlorinated wastes in soil, including wood treating preservative sludges (Hong et al. 1995). Thus, KPEG may serve a dual purpose—disposal of obsolete chlorinated pesticides formulations and cleanup of contaminated soils at deficient storage sites. However, substances that might inhibit reaction efficiency have not yet been sufficiently studied to make the technique universally usable. For example, the reaction can proceed without the use of organic solvents, but water will severely inhibit reaction rate (Brunelle and Singleton 1985). Thus, soil must be air dried, but liquid and sludge materials will be more difficult to process before treatment. Other substances including transition metals such as arsenic are substantial percentages of defunct inorganic pesticides, including lead and calcium arsenates; these may have an inhibitory effect also. Finally, EPA has warned that high clay content, acidity, and high natural organic matter may interfere with KPEG reactions, and does not recommend application of KPEG treatment to large waste volumes with concentrations of chlorinated contaminants above 5% (USEPA 1996).
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Fig. 2. Schematic for potassium polyethylene glycol ether (KPEG) degradation of chlorinated waste. Vapors emanating from the reaction drums are condensed into the water drum, and the scrubber containing sodium hypochlorite solution traps the remaining condensables. Vapors then pass through absorbent containing activated carbon and a molecular sieve. (From Taylor et al. 1990.)
VI. Options for Disposal of Pesticide Containers and Packaging Approximately 250 million pesticide containers are estimated to be used each year in the U.S. (Adebona et al. 1992). About 30% of this total has been estimated to be 50-lb bags, and presumably the remainder are a combination of plastic and metal containers. A survey in the state of Minnesota in 1990 indicated that nearly 40% of growers are burning and burying their used containers (Hansen and Palmer 1992). Such a disposal method is certain to create localized soil contamination as well as release toxic emissions. A. Recycling Options Containers have undergone an evolution in design over the past 15 yr, moving from single-trip packaging that must be disposed of by the farmer to returnable containers (Allison 1992). Furthermore, as the rates of application have significantly decreased with the new generation of herbicides and insecticides, packaging has also gotten smaller. Water-soluble packaging in some cases has eliminated the spillage and worker exposure problems. Ideally, all containers would be returned for refilling at the distributor’s or manufacturer’s facility and then returned to the grower for reuse. Returnable containers obviate the need for triple rinsing, which is the recommended minimum cleanout procedure mandated by FIFRA (Paulson 1998). In the U.S., numerous states have begun programs to collect nonrefillable containers (Hansen and Palmer 1992). Most of these containers are now made from high density polyethylene plastic. They are chipped at the various centralized collection centers and the plastic is then recycled into other products, including new containers as well as plastic benches and fence boards. Containers are not accepted unless they have been cleaned by triple rinsing. A pilot collec-
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tion and recycling study in Minnesota rejected 12% of containers in the first year of the study, but only 8% were rejected in the second year. Residue analysis of a water extract from cleaned containers that had contained several different emulsifiable concentrate (EC) herbicide formulations showed removal efficiencies ranging from 99.999% to 100% (Hansen and Palmer 1992). The results suggested farmers could efficiently clean up their containers. One study has compared a pressurized nozzle to a conventional triple rinsing (Lisowski et al. 1994). The pressurized washing device consisted of a sharptipped nozzle containing holes for flow of water. The container was pierced by the nozzle and the water turned on for as long as 40 sec. Under the experimental protocol, water was added to fill 25% of the emptied container, swirled, and then emptied. The process was repeated three more times with water analyzed after the fourth rinsing. The benchmark of adequate rinsing was to achieve 99.9999% removal of the residual pesticide. For most formulations tested, triple rinsing, but not pressure rinsing, met the benchmark of cleanliness. B. Incineration of Plastic and Paper Packaging Open burning either in pits or in piles on top of the ground is commonly practiced, especially in developing countries. Plastic jugs, even if triple rinsed, and bags contain residual amounts of pesticide. Burning will not be 100% efficient unless adequately high temperatures can be maintained. Simulated and field test burns with 50-lb insecticide bags have shown numerous PICs, including PAHs and low levels of dioxins (Adebona et al. 1992). After open burning of phoratecontaining bags, 2% of the remaining phorate was released to the air and 0.5% remained in the remaining solid residue (Oberacker et al. 1992). With atrazinecontaining bags, 13% of the remaining product was released into the air during burning, but 25% remained with the residual material. These results suggest that the temperatures of complete combustion for residual pesticides in their containers are either not reached or not maintained long enough to achieve desirable destruction efficiencies (i.e., 99.99% or greater). The British Agrochemicals Association (BAA) has designed and tested an on-farm incinerator unit that seems capable of maintaining temperatures of at least 800 °C. The unit is being recommended for well-rinsed packaging (Dyer 1998), and the design is simple enough to implement in developing countries. The incinerator consists of a steel 210-L drum outfitted with a metal wire grid placed 15 cm above the drum floor. Air holes are drilled in the sides of the drum, and temperatures as high as 950 °C can be maintained. At such temperatures, smoke production is nil and the generated combustion gases are similar to those from wood burning.
VII. Options for Management and Disposal of Wastewater Unused diluted pesticide formulations and equipment rinsewater contain high levels of pesticides that can lead to contamination of soil and eventually of water resources if not managed properly. Four basic options can be used to
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prevent contamination by wastewater (Dwinell 1992): reapplication of the rinsewater over the field as a dilute pesticide; reuse of the rinsewater as a diluent for subsequent batches of pesticides; disposal as a waste; or treatment to produce a nonpesticidal, nonhazardous material. Reapplication of rinsewater to the field is the cheapest and easiest solution. Its practical applicability is largely limited to individual farms or small-scale applicators, however. Larger commercial application services that travel from a central facility to a remote location may have more difficulty than small-scale farms with reapplication as a result of limited rinsewater availability or a contracting grower’s reluctance to risk soil compaction by repeated movement of heavy machinery across the field. However, for aerial applicators, in-flight cleaning systems allow rinsing of spray tanks and lines at the end of a flight and reapplication before leaving the field (Noyes 1992). Such systems can easily be adapted to ground sprayers. Precaution should be taken to ensure that treatment of a crop with rinsate will not cause pesticide residues to exceed the Codex Alimentarus maximum residues limits (MRL). If reapplication of rinsate is not feasible, than collection of all rinsewater and recycling it as a diluent in the next batch of spray is the second best alternative to wastewater management. Recycling depends on the construction of containment and collection pads that are built with concrete and are equipped with appropriate drains and storage tanks. If the rinsewater is segregated and identified by individual pesticide, or at least by registered crop use if in a mixture, then the EPA considers the stored material a rinse rather than a hazardous waste (Noyes 1992). Studies in Louisiana have shown that stored rinsewater that was segregated by crop use and used as a diluent for new herbicide sprays caused no phytotoxicity problems in soybeans, cotton, or rice (Rester 1987). In Illinois, studies with recycling of rinsewater containing corn and soybean herbicides concluded that adding the waste to the fresh mix at a 5% rate very slightly influenced the total active ingredient concentrations (Taylor et al. 1988). A. Treatment and Disposal Recycling of unused sprays is the best strategy for managing wastewater, but this is not universally practical, especially when a diversity of crops are grown that require pesticides neither compatible with one another nor registered for the alternative crop. However, before even considering treatment for disposal of rinsewater, growers should make every effort to recycle. Great expenses can ultimately be incurred by not recycling, including the risk of contaminating surface water and groundwater or being listed as a contaminated site should land use change. Numerous techniques have been studied to treat and dispose of pesticide wastewater, but processes for small generators have not been commercially developed and readily available. Disposal without treatment would involve collecting the rinsewater and removing it as a hazardous waste to some other location. This process would be too costly for any farmer or application business.
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Treatment before disposal alters the material to a nonpesticidal and nonhazardous state. To be practical, treatment processes must involve a technology appropriate to the pesticide applicator, be economically feasible, produce a less toxic and degradable product, and be in compliance with all regulations (Dwinell 1992). Treatment may involve physical, chemical, or biological processes, but a distinction needs to be made between physical processes that only reduce the volume or alter the physical state of the rinsewater and those which alter the chemical characteristics. Many of the treatment processes are still experimental, with few at even the pilot stage of development. The manufacturing industry has employed a diversity of biological, chemical, and physical treatment technologies (Atkins 1972), but small-scale generators have traditionally used soil disposal. Physical, chemical, and biological treatment techniques have been reviewed (Norwood 1990; Norwood and Randolph 1990) (Table 8) and assessed from the standpoint of feasibility and, to a lesser extent, of cost (Bridges and Dempsey 1988). It is beyond the scope of this analysis to discuss in detail all the myriad techniques, especially considering that many are experimental and may not be feasible for small-scale waste generators. Some techniques, however, have received both basic and applied study and hold promise for eventual adaptation. The following discussion focuses on physical entrapment methods and chemical oxidation methods (UV ozonation, Fenton’s reagent, photooxidation with TiO2). Each of these techniques, however, does not complete the treatment process but needs to be integrated with a second process most likely involving biodegradation. B. Physical Treatment 1. Evaporative Beds. Two types of physical treatment systems have been tested and employed, lined evaporation/degradation beds and granular activated sorption systems. In the U.S., physical entrapment systems based on the evapoTable 8. Technical options for disposal of pesticide-laden wastewater. Physical treatment Storage & recycling Lined evaporation beds Activated carbon adsorption Synthetic resin adsorption Peat adsorption Biomass adsorption Cyclodextrin encapsulation Reverse osmosis Solvent extraction
Chemical treatment UV ozonation Fenton reaction Solar photodecomposition Wet-air oxidation Supercritical water oxidation Chemical reduction Acid or alkaline hydrolysis Microwave plasma destruction Titanium dioxide
Source: After Norwood (1990); Norwood and Randolph (1990).
Biological treatment Activated sludges Trickling filters Biobeds Microbial enzymes Composting Phytoremediation Landfarming
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rative bed design, with or without a biodegradation process, are likely to come under the regulatory authority of RCRA if any of the disposed pesticides are listed as hazardous waste (Dwinell 1992). A facility can avoid a RCRA listing if nonhazardous pesticides are disposed, but the likelihood of this option is minimal considering that many currently registered compounds are RCRA listed. In Australia, the Code of Practice for the Storage and Use of Chemicals at Rural Workplaces (1994) allows for disposal of pesticide waste on private property under certain guidelines. For example, the disposal site should be dedicated and selected to avoid leaching and runoff and sited where there is no danger of contaminating dwellings, groundwater or surface water, and crops or livestock. The disposal pit should be on level ground with a depth between 0.5 and 1 m and lined with a plastic or clay barrier and a layer of lime. Bed disposal systems have been investigated both experimentally as well as in continuing function at several sites in the U.S. (Hall 1984; Winterlin et al. 1984; Dwinell 1992). The beds generally consist of an excavated pit with a plastic liner overlain by soil. The pits are open to the atmosphere to allow evaporation of water. The pits are covered with a translucent roof to prevent rain and snow from raising the water level. Wastewater is collected on a rinsate pad and pumped into the pit, usually from the bottom. Biodegradation has been shown to occur in the pits. Residues of pesticides in air above the pit have been shown to be minimal. Such a disposal system costs up to $50,000 (U.S. circa 1992) to construct and may have applicability for medium-scale operations. The cost seems prohibitive for individual growers. Furthermore, the systems would require RCRA permitting in the U.S. RCRA permitting of a physical treatment system could be avoided if it was considered to be closed loop (Dwinell 1992). Granular activated carbon systems qualify in that water is pumped through and the effluent is essentially free of pesticide residue. The effluent can then be used to rinse equipment or wash down the collection area. If the effluent was discharged into a sewer system or body of water, then it would probably be regulated under the U.S. Clean Water Act and require a National Pollution Discharge Elimination System (NPDES) permit (Dwinell 1992). 2. Carbon and Lignocellulosic Sorption. Carbon adsorption has been effectively used to treat wastewater in the pesticide manufacturing industry (Atkins 1972) and in spill control and cleanup (Becker and Wilson 1978). The granular activated carbon adsorption systems generally consist of a prefilter (e.g., sand) or alum flocculation chamber and a carbon filter (Becker and Wilson 1978; Dennis and Kobylinski 1983; Nye 1984; Dwinell 1992). One commercially operational system made by the Wilbur-Ellis Company in the U.S. has incorporated an oil filter and ozonation chamber before the carbon filter (Dwinell 1992). A pilot system known as the Carbolator used a suspended bed of carbon packed in floating porous polyethylene bags (Dennis and Kobylinski 1983). The advantage of a suspended bed was the avoidance of clogging and flow problems attendant with packed beds. The effluent water was directed back to the waste
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holding tank and then continuously recirculated through the carbon filters. The Carbolator was field tested with rinsewater containing malathion, propoxur, chlorpyrifos, dimethoate, diazinon, and 2,4-D and shown to be capable of removing residues to nondetectable levels. One drawback to the carbon adsorption systems is that any sludge from the flocculation step or the clogged sand filter and the carbon itself could be disposed of as solid waste rather than hazardous waste only if RCRA-listed pesticides were not treated. Nevertheless, by efficiently adsorbing pesticides from water, carbon reduces the volume of waste by several orders of magnitude. One possible solution to the carbon disposal or recycling problem, although still experimental, is the use of biologically active carbon filtration beds (Selim and Wang 1994). Such beds were shown capable of adsorbing and degrading nearly 99% of the influent atrazine. The system tested was designed for groundwater cleanup (concentrations in hundreds of parts per billion, ppb), but should be capable of being modified to handle concentrations of dilute rinsewater. Activated carbon may not be readily available in lesser-developed countries. However, wood charcoal, which is made from pyrolyzed wood, is commonly used as a fuel but also has filtration capabilities for cleaning wastewater (Agbanobi 1999) and sorptive capabilities for pesticides (Keerthinarayana et al. 1990; Seeman 1996). Combinations of mixed microbial cultures with wood charcoal in a reactor caused a 70% decrease in atrazine concentrations within 3 days (d) (Ghosh et al. 2001). Endosulfan in water was sorbed by wood charcoal, eluted with 10% methanol, and then treated with acclimated mixed microbial consortia (Sudhakar and Kikshit 2001). Ninety-six percent of the endosulfan was removed from the effluent. Innovative materials for adsorbing pesticides from wastewater have included spent mushroom compost, peat moss, acid-modified peat, and steam-exploded wood fibers (Kuo and Regan 1999; MacCarthy and Djebbar 1986; Mullins et al. 1989, 1992). Wood fibers and peat moss were proved as efficient as activated carbon (99% removal) in removing pesticide formulations diluted to 5000 mg/ L as well as water contaminated with concentrations less than 100 mg/L (Mullins et al. 1992). Before passing the diluted pesticide formulations through the sorbents, they were treated with Ca(OH)2 to deemulsify them. The retained pesticides were then transferred to composting bins for further treatment by biodegradation (Mullins et al. 1992; Berry et al. 1993; Willems et al. 1996). In a process somewhat similar to the use of lignocellulosic sorbents, a proprietary polymeric flocculent, SIROLAN CF, was mixed with sheep wool scour effluent containing sub-mg/L quantities of organophosphate and pyrethroid insecticides and nonylphenol ethoxylate surfactants (Jones and Westmoreland 1999). The flocculant reduced the pesticide and surfactant concentrations in the clarified wastewater by 99% and 95%, respectively. The flocculated material was then composted to achieve greater than 99% reduction of the pesticide concentrations (in mg/L). A schematic of the combined lignocellulosic adsorption/composting procedure is illustrated in Fig. 3. The bins (called solid-state fermentors) were supple-
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Fig. 3. Ozonation/bioreactor (left) and sorption/composter (right) schemes for cleaning wastewater. (Adapted from Hapeman-Somich 1992; Mullins et al. 1992.)
mented with organic amendments (cornmeal), crushed limestone, and microbial inocula in the form of animal manure or agricultural soils. Composting of concentrated sprays and formulations of diazinon, atrazine, and carbofuran using peat moss caused greater than 99% degradation of parent pesticide (Mullins et al. 1989; Berry et al. 1993). Leachability and bioavailability of pesticides remaining in composted material was negligible (Berry et al. 1993). Ecotoxicological studies on environmental safety of composting a diazinon formulation sorbed to peat moss showed that at least 60 d were required to eliminate earthworm sublethal toxicity (Leland et al. 2001). The proposed mechanism was a reduction in bioavailability rather than biodegradation of pesticide residue. Construction of filtration columns and bioreactors can be inexpensive; sewer
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pipe and a metal drum would be sufficient to hold the sorptive and bioreactive matrices. Ca(OH)2 and crushed limestone should be readily available to small waste generators. The composting process could require extended periods of time (up to 1.5 yr). Also, aeration is needed for efficient composting but can be achieved by mechanical turning. After the compost is finished, it could be spread on land or recycled into a new composting bin. 3. Biobeds. To capture spills during spray tank filling operations, biobeds have been developed in Sweden (Torstensson et al. 1994). The biobed consists of several meters of soil and adsorbent matrix, which may be straw and peat. The bed is seeded with pesticide-tolerant plants. Rails are laid over the bed to accommodate a tractor and sprayer. The sprayer tank is moved in place over the bed and filled up. Studies of leaching and degradation of test pesticides bentazon, chloridazon, and linuron show the bed to efficiently degrade the pesticides as well as prevent leaching below the bed (Torstensson et al. 1994). Such beds are easily constructed, but their ability to handle rinsewater as opposed to spills needs further research. Constructed wetlands may be considered a larger-scale biobed with potential to remediate wastewater generated during farming operations. Constructed wetlands are properly considered a type of phytoremediation process (discussed in detail under Section VIII, Cleanup of Contaminated Soils) because it relies on a combination of sorption in sediment and uptake/metabolism by resident plants. Studies thus far have focused on contaminated water emanating from agricultural facilities such as greenhouses or nurseries or from whole agricultural fields. Formulation residues of omethoate and parathion (µg/L concentrations) were removed with 100% efficiency by circulating water in a dual chamber containing tropical plants that were placed in a greenhouse so that it could be operated during temperate zone winters (Cheng et al. 2002). However, the system was not effective at removing the herbicides MCPA (36% efficiency) or dicamba (0% efficiency). On the other hand, a greenhouse-designed system using aquatic macrophytes and influent containing 7 mg/L atrazine delivered by subsurface flow effectively reduced atrazine residues to nondetectable levels within 1 month (McKinlay and Kasperer 1999). A constructed wetland designed to trap wastewater from a nursery operation was 24% efficient in removing atrazine within 7 d (Runes et al. 2002). A laboratory microcosm to simulate the constructed wetland at the nursery operation showed 88% removal efficiency within 56 d (Runes et al. 2001a). In summary, biobeds or constructed wetlands are comparatively simple technologies that can be designed with inexpensive materials. However, engineering expertise would be required to ensure that wastewater is fully contained within the system. Further work is needed on constructed wetlands to determine their ability for remediating higher concentrations of pesticides in rinsewater. Perhaps, coupling a biobed as described from Sweden with directed flow of effluent into a constructed wetland might be a solution for highly contaminated rinse-
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water. Constructed wetlands on small farms or, alternatively, a wetland shared by several farms would have other positive environmental benefits (for example, wildlife enhancement). C. Chemical Treatment Tested chemical treatment methods for pesticide waste degradation include photolysis, hydrolysis, dehalogenation, and oxidation (Hapeman-Somich 1992). Although all methods should still be considered experimental in that most research has been confined to the laboratory, several pilot studies have shown some feasibility for field use. Chemical methods do not completely mineralize pesticides to compounds without regulatory concern, but coupling chemical treatment with biodegradation can effectively complete degradation (Hapeman and Torrents 1998). Nevertheless, research on chemical treatment methods indicates that low-cost, locally available technologies for wastewater treatment may be feasibly developed. 1. Photolysis. Direct photolysis of organic chemicals is inefficient owing to the poor overlap between their absorption spectrum and that of sunlight. However, indirect photolysis, whereby a sensitizer molecule absorbs UV first, has been shown very effective in photodegradation of numerous organic contaminants (Watts et al. 1988). The sensitizer, for example, a dye or a humic acid, absorbs UV and is raised from the ground state to the triplet state. The energy is transferred to oxygen, generating singlet oxygen. The highly reactive singlet oxygen efficiently oxidizes a a wide variety of different contaminants. One proposal for degradation of pesticide-contaminated wastewater involves the use of solar ponds or lagoons (Watts et al. 1988). A sensitizer, for example, methylene blue or riboflavin, could be added to the water. In the presence of photosensitizers, bromacil half-life was 1 hr and 7 hr at pH 9.2 and 6.8, respectively (Acher and Saltzman 1980). Within 72 hr, atrazine (10 mg/L) was degraded by 80% in water treated with riboflavin (Cui et al. 2002). A limiting factor in a wastewater lagoon would be the rate of aeration. Optimal oxygen requirements for indirect photolysis have been studied for bromacil and terbacil to help design an aeration system (Watts et al. 1988). Another problem with a UV-driven system is that formulations and rinsates will be turbid as well as have a mixture of chemicals that could quench the photodegradation reaction (Hapeman-Somich 1992). However, DDT (µg/L levels) dechlorination in aqueous phase with intense UV illumination was enhanced by the presence of surfactant (Chu 1999). The half-life for DDT dechlorination was less than 4 min, but pertinently, DDE and DDD were also dechlorinated after their formation. Research of waste lagoons containing pesticide rinsates should be conducted to determine degradation rates under natural sunlight. 2. Hydrolysis. Under laboratory conditions, ester linkages in numerous pesticides, including pyrethroids, carbamates, organophosphates, and acetanilides,
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can be hydrolyzed in solutions of elevated pH levels. Under field conditions, however, formulation ingredients of commercial products may act as buffering agents and inhibit reactivity (Hapeman-Somich 1992). Calcium hydroxide, viewed as a safer alternative to sodium and potassium hydroxide, hydrolyzed a fenitrothion formulation diluted to 10 mg/L with a half-life of 69 min at room temperature (Desmarchelier 1987). Hydrolysis of fenitrothion by peroxide (0.1–0.2 M) in the presence of 1% w/v sodium carbonate occurred with a half-life 19 min (Desmarchelier 1987). Under basic conditions, sodium perborate liberates peroxide anion that is more reactive to OP esters than hydroxyl ion (Lee et al. 1984). Sodium perborate enhanced the hydrolysis of mevinphos, diazinon, methyl parathion, malathion, and parathion in lake water (10 mg/L) at pH 9.88 (Qian et al. 1985). Half-lives were generally 20 min or less. When soil was present (up to 5% w/v), the reaction was significantly slowed as a result of the perborate reacting with the soil. However, a fourfold increase from 0.03 to 0.12 M perborate enhanced hydrolysis. An estimation of cost of treating a 40,000-L wastewater pond contaminated with organophosphorus (OP) insecticides indicated the method is comparatively cheap (US$142 in 1984) and thus may have applicability to small farms that use a lot of OP insecticides (Qian et al. 1985). Divalent metal ions and metal oxides have also been shown to catalyze hydrolysis of OP insecticides. Copper (+2) was the most effective metal for catalysis of five phophorothionate insecticides (chlorpyrifos-methyl, zinophos, diazinon, parathion-methyl, and ronnel) and two phosphorooxonates (chlorpyrifosmethyl oxon and paraoxon) (Smolen and Stone 1997). The optimal pH was about 5, and the reaction was inhibited at higher pH levels as a result of reduction in copper solubility. One drawback to catalysis by metal oxides could be the formation of products that may retain signficant toxicity. Hydrolysis of demeton, disulfoton, and thiometon in the presence of three iron oxides (hematite, goethite, and ferrihydrite) at pH 8.5 and deaerated conditions resulted in the formation of the persistent compound 1,2-bis(ethylthio)ethane (Dannenberg and Pehkonen 1998). 3. Dehalogenation. The KPEG process can dehalogenate recaclitrant compounds such as PCBs and dioxin congeners, and it has been shown effective for stored drums of phenoxy herbicides, including 2,4,5-T. However, the reactionquenching effects of water make KPEG inapplicable to wastewater unless the water is exchanged with an organic solvent. The use of zero-valent iron to reductively dechlorinate small halogenated aliphatics is a recent development in dehalogenation chemistry that seems promising for contaminated wastewater. The dehalogenating abilities of metallic iron were discovered serendipitously during background research surrounding the remediation of groundwater contaminated by chlorinated solvents (Rock et al. 1998). Subsequently, zero-valent iron was reported to catalyze the reduction of nitroaromatics to anilines (Agrawal and Tratnyek 1996), which are not less toxic but are more readily biodegradable than the parent compound (Weber 1996).
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Most of the research with zero-valent compounds has focused on chlorinated methanes and ethanes. The reaction mechanism is thought to proceed essentially as an iron corrosion process with electrons transferred to a chlorinated organic compound (Eq. 5) rather than oxygen (Eq. 4) (Matheson and Tratnyek 1994): 2Fe0 + O2 + 2H2O → 2Fe2+ + 4OH− +
2Fe + RX + H → 2Fe + RH + X 0
2+
−
(4) (5)
The most favored hypothesis to explain the reaction mechanism is that the electron transfers take place on the iron surface (Matheson and Tratnyek 1994). Thus, surface area and mixing (i.e., increasing diffusion from the aqueous to the solid phase) are important factors that can control reaction rate. Water solubility limitations and competitive sorption processes that limit approach to the iron surface potentially may be overcome with the use of electron mediators such as metal-containing prophyrins, quinones, and natural organic matter (Weber 1996). These molecules would serve to transfer electrons to the contaminant remotely from the surface. The experimental designs for testing reductive dehalogenation and denitrification require an aerobic environment. Thus far, field testing has been conducted for cleanup of contaminated groundwater of chlorinated hydrocarbon solvents. Pesticides found capable of being dechlorinated under laboratory conditions include DDT (Sayles et al. 1997), alachlor, metolachlor (Eykholt and Davenport 1998; Comfort et al. 2001), EDB (Rajagopal and Burris 1999), atrazine (Dombek et al. 2001), benomyl, picloram, and dicamba (Ghauch 2001). Parathion was reduced to amino parathion (Agrawal and Tratnyek 1996). Zero-valent iron holds great promise for in situ remediation of contaminated groundwater. However, it requires engineering and groundwater hydrology in the field. From a practical perspective, the present field applicability is limited to in situ remediation of groundwater contaminated with chlorinated solvents. 4. Oxidation. Compounds most susceptible to oxidation contain heteroatoms with lone pairs of electrons (oxygen, nitrogen, sulfur), and thus pesticides should be very susceptible to oxidative treatment. Two main oxidation processes have been studied, direct oxidation with hydrogen peroxide and ozone, and advanced oxidation processes, where oxidants are combined with UV light, iron salts, or titanium dioxide to generate free hydroxyl radicals (Hapeman and Torrents 1998). Advanced oxidation processes for pesticide waste treatment are discussed. Ozonation. The pesticide manufacturing industry (Atkins 1972) and municipal wastewater treatment plants (Masten and Davies 1994) have used ozone as an oxidant for degrading organic contaminants in wastewater. Early studies focused on drinking water contaminated with chlorinated hydrocarbon pesticides at their maximum limits of solubility. Ozone alone (3.9% by volume) or in combination with aeration was shown to effectively remove 80%–90% of the lindane, aldrin,
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and dieldrin present in carbon filtered water or river water (Buescher et al. 1964). Ozone is a strong oxidant but its reaction alone with organic compounds can be slow. Ozone in combination with UV, or under basic conditions in combination with H2O2, ultimately produces hydroxyl radicals that can significantly transform many types of organic compounds within minutes (Table 9). Early studies with pesticide wastewater used ozone and UV in a mobile pilot reactor to successfully degrade paraquat, 2,4-D, and atrazine (Kearney et al. 1984). Ozonation alone and ozone/UV catalyzed degradation of pesticides with equal efficiency (Kearney et al. 1987), but real wastewater tends to be opaque, rendering UV ineffective (Hapeman-Somich 1992). For any wastewater treatment process, knowledge of the bioactivity and ultimate fate of degradation products provides enhanced safety. Studies with atrazine, alachlor, paraquat, and mixed herbicide wastes have followed reaction pathways during ozonation alone and with UV irradiation (Kearney et al. 1985, 1988; Somich et al. 1988). Depending on the herbicide, typical transformations include ring opening, rearrangements, dealkylations, and additions of oxygen. Mineralization, however, could be achieved by passing the ozonation products through soil columns, continuously stirred bioreactors, or fixed-film bioreactors (Somich et al. 1990; Leeson et al. 1993). Such final disposition has been studied with atrazine (Leeson et al. 1993; Hapeman et al. 1995), alachlor (Somich et al. 1988), and bromacil (Acher et al. 1994). Coumaphos cattle vat dip waste was not degraded efficiently by UV ozonation (Kearney et al. 1986). However, destruction of the waste (1500 mg/L) was achieved by first treating it with a culture of Flavobacterium sp. Radiolabeled coumaphos was hydrolyzed to the toxic metabolite chlorferon. Transfer of the waste to a UV-ozonation chamber for several hours degraded all the chlorferon. The resulting mixture was then incubated in soil, and more than 70% of the radiolabel was released as CO2, compared to none released with no waste treatment and only 20% released following the ozonation or bacterial treatment alone. Whether in combination with UV or H2O2, it is now clear that ozonation only partially transforms pesticides (Hapeman-Somich et al. 1992), and coupling of chemical treatment with biotreatment is necessary to more fully degrade the pesticide waste and dispose of metabolites that may still possess unwanted biological activity. A scheme for small-scale field disposal by a two-stage ozonation-biomineralization process has been published (Hapeman-Somich 1992) (see Fig. 3). The major investment in cost for such a system would be the ozone generator and pumps; however, liquids could be moved from one part of the system to the other by use of gravity flow. The containers could be made of 55gal drums. The only reagents required would be hydrogen peroxide and sodium carbonate, which should be readily available to small waste generators. Nutrient and organic amendments for the bioreactor, which could simply consist of locally collected soil, can be obtained by addition of food processing waste, cereal meal, or some other readily available plant source. The microbially rich environ-
154
Table 9. DT50% (time to 50% dissipation) of pesticides subjected to ozonation alone, UV ozonation, or UV-H2O2. Pesticide (oxidation process)
C0 (mg/L)
Pesticide form or source
DT50% (min)
Reference
10–1000
Diluted commercial formulation
⬃5 to ⬃50
Kearney et al. 1987
Alachlor (UV-O3)
10–1000
Diluted commercial formulation
<10 to ⬃25
Kearney et al. 1987
240
Analytical grade Diluted commercial formulation
8.6 1.9
Somich et al. 1988
⬃25
Somich et al. 1990
<2
Beltran et al. 1994
Alachlor (O3) Atrazine (H2O2-O3)
Not specified
Pesticide waste rinsate
Atrazine (UV-O3)
10.8
Atrazine (UV-O3)
10–1000
Diluted commercial formulation
<10 to 60
Kearney et al. 1987
33 100
Analytical grade Diluted commercial formulation
<5 16
Kearney et al. 1988
10–1000
Diluted commercial formulation
<25 to >360
Kearney et al. 1987
⬃3
Torrents et al. 1998
⬃10 to ⬃285
Kearney et al. 1987
Atrazine (O3) Bentazon (UV-O3) Bromacil (O3) Carbofuran (UV-O3)
100 10–1000
Analytical grade
Purified from formulation Diluted commercial formulation
Carbofuran (O3)
100
Technical grade (pH 2 water)
16.5
Carbofuran (UV-O3)
100
Technical grade (pH 2 water)
<5
Carbofuran (UV-O3)
100
Technical grade (pH 6.5, 9.0)
12, 16
Coumaphos (UV-O3)
1500
Cattle dip waste
>360
Benitez et al. 2002 Benitez et al. 2002 Kuo 2002 Kearney et al. 1986
A.S. Felsot, K. Racke, and D. Hamilton
2,4-D isooctyl (UV-O3)
Table 9. (Continued). Pesticide (oxidation process) Cyanazine (H2O2-O3) Cyanazine (UV-O3) Imazaquin (H2O2-O3)
C0 (mg/L) Not specified 10–1000 57–893
Pesticide form or source Pesticide waste rinsate Diluted commercial formulation Technical grade
DT50% (min) <10 min <10 to ⬃60 ⬃1 to 7 ⬃5
Reference Somich et al. 1990 Kearney et al. 1987 Massey and Lavy 1997
10
Diluted commercial formulation
1 45
Purified technical grade
9; 10.4 4.2; 4.2
Benoit-Guyod et al. 1986
Technical grade (pH 6.5, 9.0)
114, 64
Kuo 2002
Mevinphos (UV-O3) Metolachlor (H2O2-O3)
100 Not specified
Pesticide waste rinsate
<20
Metolachlor (UV-O3)
10–1000
Diluted commercial formulation
⬃5 to ⬃20
Metribuzin (UV-O3)
10–1000
Diluted commercial formulation
⬃5 to ⬃270
Paraquat (H2O2-O3)
Not specified
Paraquat (UV-O3)
Trifluralin (UV-O3)
1500 1500 100 10–1000
Pesticide waste rinsate Formulation, acetone as sensitizer Formulation, no acetone Formulation, acetone Diluted commercial formulation
Kearney et al. 1987
Somich et al. 1990 Kearney et al. 1987
Pesticide Waste
Malathion (UV-O3) MCPA (UV-O3; O3)
Kearney et al. 1987
⬃45
Somich et al. 1990
⬃360 >480 ⬃10
Kearney et al. 1985
⬃20 to ⬃210
Kearney et al. 1987
Source: Dissipation times are extrapolated from either graphical representations in the reference or calculated assuming first order half-life.
155
156
A.S. Felsot, K. Racke, and D. Hamilton
ment of the soil bioreactor would enhance mineralization of the ozonated wastewater or transformations leading to bound residues (Somich et al. 1988, 1990); thus, leachate would be comparatively clean, and the bioreactor matrix could be recycled or disposed of by landfarming (discussed in Section VIII). Although a pilot-scale mobile UV-ozonation reactor has been tested with some success on pesticide rinsewater in the field (Kearney et al. 1984), the general applicability of the process remains questionable because of the potential for UV and OH quenching from formulated materials. Furthermore, more research is needed to perfect integration with a biodegradation technique to ensure total detoxification of transformation products. Such a system is not feasible for individual farms but could be developed for a single location serving many small waste generators. Fenton’s Reagent. Hydrogen peroxide in combination with a ferrous salt, known as Fenton’s reagent, functions as an oxidizing reagent through the formation of hydroxyl radical at acid pH (Eq. 6). Fe2+ + H2O2 → Fe3+ + HO− + OHⴢ
(6)
Hydroxy radical may be scavenged by reaction with ferrous ion (Eq. 7), or it may react with an organic compound (Eq. 8). OHⴢ + Fe2+ → HO− + Fe3+
(7)
OHⴢ + RH → ROH
(8)
Studies with industrial effluents containing phenol and substituted phenols showed significant oxidation of phenolic constituents (Eisenhauer 1964) in the presence of ferrous ammonium sulfate and hydrogen peroxide. The reaction rate was comparatively slower with more substituted phenols and was nil with pentachlorophenol. Ferric ions in the presence of artificial UV or sunlight are photoreduced to hydroxy radicals and consequently the formation of Fe(OH)2+ complex, especially at acidic pHs. When ferric perchlorate, Fe(ClO4)3ⴢH2O (0.26 mM) was added to a solution of atrazine (5.2 µM, 1.2 mg/L) and exposed to a mercury arc lamp, atrazine half-life was less than 2 min compared to 1500 min in the absence of the iron salt (Larson et al. 1991). The reaction was somewhat slower when ferric sulfate (0.21 mM) was substituted (24 min compared to 25 hr without iron salt). Natural (i.e., nondistilled) water inhibited the reaction with ferric perchlorate, as evidenced by an increase in half-life to 900 min. During the 1990s, numerous technical grade pesticides were tested for degradation by both the dark and photoassisted Fenton reaction (Table 10). The dark reaction proceeds as characterized by reactions 6 through 8 (Eqs. 6–8), and the photoassisted Fenton reaction can be described by Eq. 9 and 10 (Huston and Pignatello 1999): Fe2+ + H2O2 → Fe(OH)2+ + OH
(9)
Fe(OH)2+ + hv → Fe2+ + OH
(10)
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157
With few exceptions, half-lives are less than 5 min. Furthermore, mineralization of a variety of pesticides have been observed (Pignatello 1992; Sun and Pignatello 1992, 1993a,b; Huston and Pignatello 1999), including some PAHs in creosote (Engwall et al. 1999). The photoassisted Fenton reaction may be nearly as efficient on formulated pesticides as on technical grade active ingredients (Table 11) (Huston and Pignatello 1999). Polymer-based controlled-release formulations may inhibit the reaction owing to mass transfer limitations, but greater efficiency can be achieved by raising the temperature of the reaction (Table 11) (Huston and Pignatello 1999). The Fenton reaction tends to run at acid pHs, which can pose a safety problem, especially if the reaction mixture must be discharged to the environment. Use of chelators, including picolinic acid, gallic acid, and rhodizonic acid, allowed both the dark and photoassisted reaction to proceed at pH 6 (Sun and Pignatello 1992). An electrochemical Fenton treatment (EFT) system was developed that would allow flowthrough of contaminated solution and a reaction at circumneutral pH (Roe and Lemley 1997; Pratap and Lemley 1998). A current is applied across sacrificial iron electrodes immersed in aqueous sodium chloride to generate ferrous ion. Hydroxyl radical is generated by the addition of hydrogen peroxide. The system was first demonstrated to degrade the herbicides alachlor, metolachlor, atrazine, cyanazine, and picloram with 50% disappearance times of less than 30 min (Pratap and Lemley 1994). The EFT system has several disadvantages compared to the conventional Fenton treatment system: slower degradation rate; potential sorption of target compounds and degradation products to iron solids; and inhibition of reaction owing to hydroxide ion production at the cathode and circumneutral pH (Saltmiras and Lemley 2000). The limitations of EFT were overcome by the development of an anodic Fenton treatment (AFT) system in which a salt bridge was used to connect two half-cells. Pesticide-contaminated water is degraded in the anode half-cell, which has a sacrificial iron electrode that generates ferrous ions under a low current of 0.12 A. The cathode is graphite and is held in a deionized water half-cell. Initial studies showed that ethylene thiourea (ETU), a degradation product of EBDC (ethylene bisdithiocarbamate) fungicides, is transformed to undectable levels within about 5 min in a zero-order reaction (Saltmiras and Lemley 2000). The AFT process was further optimized and applied to the degradation of 2,4-D (Wang and Lemley 2001). Finally, the salt bridge was replaced with an ion-exchange membrane, and a single system was robust enough to be repeatedly used 100 times (Wang and Lemley 2002). In the membrane AFT, carbaryl was completely transformed within 5 min to noninsecticidal degradation products. The potentially bioactive products resulting from the dark-assisted Fenton reaction with atrazine were further mineralized by cultures of Rodococcus corallinus and Pseudomonas sp., releasing 70% of the ring labeled carbon as CO2 (Arnold et al. 1996). The combination of the Fenton reaction and microbial biodegradation was tested on a mixed pesticide rinsate containing atrazine, cyanazine, metolachlor, alachlor, and EPTC.
158
Table 10. Time to 50% dissipation (DT50%) of pesticides by dark or photoassisted (UV) Fenton’s reagent. C0 (mg/L)
DT50% (min)
Reaction conditions
2,4,5-T 2,4,5-T 2,4-D 2,4-D Alachlor Aldicarb Atrazine Atrazine Atrazine Atrazine Azinphos-methyl Captan Carbaryl Carbofuran Creosote (PAHs; phenolics) Diazinon Dicamba EPN Disulfoton
25.5 25.5 22 22 27 19 30 21.6 21.6 30 31.7 30.1 20.1 22.1 0.019–16.7
<5 <5 <5 <5 <10 <2 <5 <5 <10 <1 <2 <10 <2 <10 <5
0.001 M Fe3+; 0.01 M H2O2; UV; pH 2.8 (NaClO4) 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; 0.001 M Fe3+; 0.01 M H2O2; UV-VIS; pH 2.8 (H2SO4) 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 Electrochemical system; 0.0036 M Fe+2; 0.025 M H2O2 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.027 Fe2+; 0.027 H2O2; pH 3 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.001 M Fe3+; 0.01 M H2O2; UV; pH 2.75
10 22.1 10 27.4
63 <10 53.3 <10
0.00005 0.00005 0.00005 0.00005
M M M M
Fe2+; Fe3+; Fe2+; Fe3+;
UV; 0.29 M H2O2 0.01 M H2O2; UV; pH 2.8 UV; 0.29 M H2O2 0.01 M H2O2; UV; pH 2.8
Reference
pH 6 pH 6
pH 6
pH 6
Pignatello 1992 Sun and Pignatello 1993b Pignatello 1992 Sun and Pignatello 1993b Huston and Pignatello 1999 Huston and Pignatello 1999 Pratap and Lemley 1998 Sun and Pignatello 1993b Huston and Pignatello 1999 Arnold et al. 1995 Huston and Pignatello 1999 Huston and Pignatello 1999 Sun and Pignatello 1993b Huston and Pignatello 1999 Engwall et al. 1999 Doong and Chang 1998 Huston and Pignatello 1999 Doong and Chang 1998 Huston and Pignatello 1999
A.S. Felsot, K. Racke, and D. Hamilton
Pesticide
Table 10. (Continued). Pesticide
DT50% (min)
Reaction conditions
30 33 145 10 34.6 400 10 12 50 52.6 28.4 28.4 55 7.95 10 24.1 24.1 20.9 20.2 33.5
<5 <10 ⬃20 57.8 <30 87 33 <5 ⬃5 ⬃2 ⬃2 <10 <5 <5 34.7 <10 <5 <5 <2 >300
Electrochemical system; 0.03 M H2O2; stepwise addition 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.0018 M Fe2+; H2O2 0.22 M; Cu2+ enhanced 0.00005 M Fe2+; UV; 0.29 M H2O2 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.0018 M Fe2+; H2O2 0.22 M; Cu2+ enhanced 0.00005 M Fe2+; UV; 0.29 M H2O2 Electrochemical system; 0.03 M H2O2; stepwise addition Fe2+ 0.0018 M; H2O2 0.22 M; Cu2+ enhanced 0.001 M Fe3+; UV, 0.01 M H2O2 0.001 M Fe3+; UV, 0.01 M H2O2 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 Electrochemical system; 0.0036 M Fe+2; 0.025 M H2O2 1 mM Fe3+; 10 mM H2O2; UV; pH 2.75 0.00005 M Fe2+; UV; 0.29 M H2O2 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; pH 6 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; pH 6 0.00005 M Fe3+; 0.01 M H2O2; UV; pH 2.8 0.001 M Fe3+; 0.01 M H2O2; UV; chelator 0.001–0.1 M; pH 6
Reference Roe and Lemley 1997 Huston and Pignatello 1999 Dowling and Lemley 1995 Doong and Chang 1998 Huston and Pignatello 1999 Dowling and Lemley 1995 Doong and Chang 1998 Roe and Lemley 1997 Dowling and Lemley 1995 Pignatello and Sun 1995 Pignatello and Sun 1995 Huston and Pignatello 1999 Pratap and Lemley 1998 Engwall et al. 1999 Doong and Chang 1998 Huston and Pignatello 1999 Sun and Pignatello 1993b Sun and Pignatello 1993b Huston and Pignatello 1999 Sun and Pignatello 1993b
Pesticide Waste
Malathion Malathion Malathion Malathion Methoxychlor Methamidophos Methamidophos Methyl parathion Methyl parathion Methyl parathion Metolachlor Metolachlor Metolachlor PCP Phorate Picloram Picloram Propoxur Simazine Trifluralin
C0 (mg/L)
159
160
A.S. Felsot, K. Racke, and D. Hamilton
Table 11. Approximate time to 50% dissipation (DT50%, min) of technical grade and formulated carbofuran and alachlor treated with the photo-Fenton process. Formulation Pesticide AI Carbofuran @ 25 °C Alachlor @ 25 °C Alachlor @ 40 °C Alachlor @ 50 °C
Technical
Flowable
Emulsifiable concentrate
MicroTech
5 2 — —
5 — — —
— 17 — —
— 110 70 35
Source: Modified from Huston and Pignatello (1999).
Recent work has shown that microorganisms generating H2O2 can drive the Fenton reaction (McKinzi and Dichristina 1999). Pentachlorophenol was incubated with a culture of the facultative anaerobic bacterium Shewanella putrefaciens strain 200 in the presence of ferric salts. The bacterially produced H2O2 resulted in oxidation of the ferrous ions to ferric ions and consequent generation of hydroxyl radical. The advantage of such a system is its operation at neutral pH and no requirements for exogenous H2O2 or UV irradiation. Photocatalysis with Titanium Dioxide. Heterogenous photocatalysis on titanium dioxide has been discovered to oxidize quickly a very wide variety of organic contaminants as well as to reduce heavy metals. TiO2 is a semiconductor, and when it is irradiated an electron is promoted from a valence band to a conduction band (Muszkat 1998). The valence band is essentially a positive hole that is capable of oxidizing adsorbed electron donors such as H2O or OH, generating hydroxyl radicals. The hydroxyl radicals can react quickly with organic species in solution. Photocatalytic oxidation of carbaryl resulted in complete mineralization to CO2 within 1 hr after initiation of the reaction (Pramauro et al. 1997). The reaction was not inhibited by surfactants, suggesting that pesticides in rinsates can be degraded by TiO2 (Prevot et al. 1999). Other pesticides studied in experimental treatment systems using photocatalysis include atrazine (Pelizzetti et al. 1990), alachlor (Penuela and Barcelo 1996), 2,4-D (Lu and Chen 1997), monocrotohos (Ku and Jung 1998), monuron (Pramauro et al. 1993), prometyrn and prometon (Borio et al. 1998), pentachlorophenol (Mills and Hoffmann 1993), propoxur (Lu et al. 1999a), dicamba (Bianco-Prevot et al. 2001), and pentachlorophenol (Pecchi et al. 2001). The broad capability of TiO2-catalyzed photodegradation and speed of reaction was illustrated in one study of four triazine herbicides (atrazine, propazine, cyanazine, prometryne) and four OP insecticides (parathion, methyl parathion, bromophos, and diclofenthion) (Konstantinou et al. 2001). Half-lives for transformations leading to detoxification ranged from 10 to 38 min. Although most experimental systems employed suspensions of TiO2 alone,
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161
recent research has investigated the effect of H2O2 and an immobilized catalyst. In a system of suspended TiO2 supplemented with H2O2, the photocatalytic halflife of parathion decreased more than twofold (from 99 to 40 min) compared to a system with TiO2 alone (Chen et al. 1998). Hydrogen peroxide has been added to enhance oxidation of prometryn, prometon, and propoxur in the presence of TiO2 immobilized on membranes or adsorbents (Borio et al. 1998; Lu et al. 1999b). Cyanuric acid can form as a stable final photoproduct in systems degrading trazine herbicides. Thus, some wastewaters might need further treatment before discharge to the environment. However, pretreatment of formulations of asulam and fenitrothion in water with ozone before photocatalysis with TiO2 enhanced degradative loss of parent compound and degradation products as measured by total organic carbon content (Tanaka et al. 1992). Use of photocatalysis may play a role in future pesticide wastewater disposal, but at present little research has been conducted with pilot systems for fieldgenerated rinsates. Mixed results were obtained when five pesticide formulations were diluted, mixed with a small amount of soil, and irradiated in the presence of TiO2 catalyst in a laboratory apparatus (Chiarenzelli et al. 1995). Propyzamide and chlorpyrifos were somewhat recalcitrant to degradation after 48 hr, but dicloran and triadimefon significantly degraded within 24 hr. Permethrin in a diluted formulation with soil particles was much more resistant to degradation than the technical grade compound in water. In another experiment oriented toward making TiO2-catalyzed photolysis practical for wastewater, carbaryl was subjected to a flowthrough system (PerisCardells et al. 1993). The half-life was 26 sec with 99% destruction in 1 min of irradiation. The concentration was only 0.2 mg/L, so applications to wastewater remain untested. Furthermore, research has not sufficiently investigated how formulation surfactants and soil particles would affect oxidation rates. Most research with TiO2 catalyst has involved laboratory-scale systems employing high-intensity artificial UV sources. Such materials may be difficult to acquire and power in lesser-developed countries. However, a commercial-scale flowthrough system using solar energy has been developed for degradation of wastewater (Malato et al. 2002). This system consists basically of looped transparent tubing on an aluminum metal solar panel. The wastewater is recirculated through the loops while solar energy is transferred from the collector into the tubing. Pilot-scale versions have successfully degraded within minutes a wide range of pesticide formulations diluted in water (Marques et al. 1996; Minero et al. 1996; Herrmann et al. 1998; Malato et al. 2000). The rate of pesticide degradation can be enhanced by addition of sodium peroxydisulphate (Minero et al. 1996). Measurements of total organic carbon in addition to parent pesticide in the pilot systems proved that degradation products were also being efficiently degraded, although at a slower rate.
VIII. Cleanup of Contaminated Soils Because of the complexity and diverse characteristics of soil matrices, as well as the extreme variability in waste concentrations over an area, cleanup of soil is more difficult than handling concentrated product formulations or wastewater.
162
A.S. Felsot, K. Racke, and D. Hamilton
Obviously the most desirable waste management technique would be to prevent soils from becoming contaminated in the first place with appropriate product and wastewater stewardship. Despite implementation of best management practices, soils do become contaminated. Cleanup of soils needs immediate as well as vigilant attention as the comparatively high concentrations and resulting prolonged persistence increase the potential for leaching and runoff. Two general options exist for cleanup of contaminated soils—separation of the chemical from the soil and direct treatment of the soil. Either option most often requires excavation and physical removal of the soil, but methods for in situ remediation are undergoing study. The direct methods can be categorized as either physical and chemical destruction or biological transformation and degradation. Separation technologies and ex situ and in situ direct soil treatment techniques are briefly discussed with a perspective on their feasibility for implementation by small waste generators. A. Soil Separation Techniques Soil separation techniques remove contaminants from the soil and concentrate them for further treatment. In general, the contaminants are neither degraded nor detoxified. Known techniques are physical in nature and can be applied in situ or ex situ (Koustas and Fischer 1998). In situ techniques include radiofrequency heating, electrokinetic treatment, and soil flushing. Ex situ techniques include soil washing, solvent extraction, and thermal desorption. Radiofrequency (RF) heating involves placing electrodes in boreholes in the soil. When a radiofrequency band is applied, molecules vibrate sufficiently to heat the soil. Both water and contaminants can be heated sufficiently to volatilize and be recovered in a vacuum at the surface. Application of the method to pesticide contaminated Superfund sites in the U.S. have shown removal of chlorinated cyclodiene pesticides with efficiencies of 97%–99.9% (Koustas and Fischer 1998). Electrokinetic (EK) treatment is somewhat similar to RF heating but relies instead on setting up a voltage in the soil by placing a cathode in one hole and the anode in another (Rock et al. 1998). The objective is to make the contaminants move from the vicinity of one electrode to the other where they can be collected. The mechanism of movement is believed to be electroosmosis, that is, mass transfer in moving water. Both RF heating and EK treatment are still too experimental to be of use to small waste generators. Even though RF heating has been applied successfully in pilot studies, it requires too much engineering expertise and monitoring to be of practical use. Soil flushing and soil washing recover contaminants from soil in situ and ex situ, respectively, using water as the solvent medium (Koustas and Fischer 1998; Rock et al. 1998). Solvent extraction is an ex situ technique that employs either an organic solvent or a supercritical fluid. Subcritical water, which is hot water under just enough pressure to maintain the liquid state, has been used successfully on both the laboratory (8 g soil) and pilot scales (8 kg soil) to remove a mixture of herbicides from a historically contaminated soil (Lagadec et al.
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2000). At 250 °C and 50 bar pressure, pendimethalin at 415 mg/kg was reduced to nondetectable levels (<0.1 mg/kg) in 2 hr. A chemical leaching procedure has been tested in Australia for decontaminating soil at old cattle dipping sites (Van Zwieten et al. 1998). The proprietary process, known as Geo2, involves several different chemical leaching solutions. More than 90% of the arsenic and DDT contamination was removed (initial concentration in soil, 2640 mg/kg and 4521 mg/kg, respectively). The arsenic could be recovered from the leaching solution, but the DDT was chemically degraded during the leaching process. The soil required further investigation to determine its suitability for replacement at the original waste site. Analogously to washing techniques, soil can be heated in an enclosed chamber and contaminants desorbed by vaporization followed by trapping out of the exhaust gas or destruction in an afterburner (Troxler 1998). DDT in contaminated soil excavated at a cattle dipping site in Australia was successfully removed by a thermal desorption process employing temperatures of 450–500 °C (Van Zwieten et al. 1998). The exhaust gas containing vaporized DDT was conducted to a hot gas filtration unit, followed by an afterburner where the residual vaporized DDT was thermally degraded at 1100 °C. Soil washing and thermal separation techniques are impractical for small waste generators owing to the cost of solvents and the equipment required to remove and then treat the recovered waste. Techniques relying on aqueous extraction may face limitations because of the wide solubility differences between pesticides likely to occur as soil contaminants in addition to problems with nonequilibrium sorption. Thermal desorption relies on equipment similar to incinerators, making the technology impractical for developing countries. Furthermore, all the techniques have had limited field testing other than at Superfund sites. Despite the lack of practicality of soil separation techniques for small waste generators, the principles can be applied to remediating contaminated groundwater. The “pump and treat” approach is essentially a soil washing technique. Continued pumping of a contaminated well into a collection area allows recharging of the aquifer. Eventually the contaminants are diluted to nondetectable levels or at least to levels not considered hazardous. The collected water can be sprayed over crop land to effect normal biodegradation in the soil. Such a method may be feasible for small waste generators in lieu of drilling a new well. B. Physical and Chemical Treatment 1. Incineration. Combustion in a rotary kiln incinerator is the method most commonly used for destruction of hazardous waste, especially in soils. As with incineration of obsolete stocks, emission of toxic PICs and unburned residual material is of concern, although incinerated soil could be returned to its site of removal if heavy metal concentrations were not a problem. Incineration of soil has recently been reviewed along with a presentation of a case study about
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pentachlorphenol- and creosote-contaminated soil (about 90,700 tonnes) from a sawmill and wood treatment plant in the U.S. (Steverson 1998). The incineration process in a rotary kiln physically occurs in two chambers. The soil is fed into a primary chamber where high temperatures (≥540 °C) desorb the pesticide into a flowing gas stream (Steverson 1998). To improve desorption, soil may have to be screened to remove rocks and lumps or it may have to be blended to reduce moisture or clay content. The gas is directed into a secondary chamber where the desorbed pesticide is combusted at temperatures of 1000–1200 °C. The hot gas stream is cooled as resulting gases and PICs are removed by a series of filters, sorption beds, and scrubbers. Incineration of soil is probably the first choice for industry facing the prospects of large-scale cleanup of contaminated soil. For smaller agrichemical application businesses and farms with large amounts of contaminated soil, incineration will not be economically feasible. And, in developing countries, incinerators for contaminated soil and obsolete pesticide stocks are unavailable. Moreover, programs to help developing countries dispose of obsolete stocks have not addressed site contamination. Thus, research on soil remediation adaptable to small-scale operations is urgently needed to prevent further environmental deterioration from deficient waste-handling procedures. 2. KPEG. Traditionally, contaminated soils have been either landfilled or, in the case of hazardous waste, incinerated. Soils contaminated with the banned organochlorine and cyclodiene incesticides should be amenable to the KPEG process (Taylor et al. 1990). Indeed, KPEG has been most often used to remediate halogenated waste such as PCBs in soil rather than the formulated chemicals themselves. The disadvantage of KPEG over rotary kiln incineration is the capacity and the fact that KPEG does not achieve destruction efficiencies greater than 99%. Thus, the residual soil will still contain both parent chemical as well as lesser chlorinated degradates. Furthermore, KPEG treatment should still be considered experimental because many chlorinated cyclodienes and hydrocarbons that are likely to be a problem in developing countries have not been tested. Nevertheless, KPEG technology is quite simple compared to incineration and seems adaptable to countries with few resources. Dechlorination reactions can make the residual chemicals in the soil more amenable to biodegradation reactions, thereby allowing land application of the treated waste. 3. Zero-Valent Iron. Dehalogenation by treatment with zero-valent iron has almost exclusively been applied to liquid waste or in situ decontamination of ground water. One pilot field study, however, has applied the technique to soil highly contaminated (>1400 mg/kg) with the herbicide metolachlor (Comfort et al. 2001). The contaminated soil was excavated, formed into windrows, and wetted periodically along with additions of zero-valent iron, aluminum sulfate, and acetic acid. After 90 d, 70% of the metolachlor was dechlorinated when only iron was used. However, when the sulfate and acetate reagents were added, transformation was more than 99% efficient. Laboratory studies, as well as ear-
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lier studies with atrazine (Singh et al. 1998), showed that transformation products resulting from zero-valent iron-mediated dehaolgenation are more amenable to biodegradation. 4. Miscellanous Chemical Treatments. Two additional chemical treatment processes are noteworthy because they may be relatively inexpensive. Chloroacetanilide herbicides (e.g., alachlor and metolchlor) are common contaminants at agrochemical application facilities, and past handling practices have resulted in contaminated soils (Felsot 1998a). Research with soil columns has shown that residues of herbicides in this chemical group can be quickly dehalogenated by amendments of the fertilizer ammonium thiosulfate and thereby prevented from leaching through the soil (Gan et al. 2002). The dehalogenation reaction resulted in loss of microbial toxicity in the Microtox assay and mutagenicity in the Ames test. Although photocatalysis using TiO2 has been almost exclusively applied to wastewater, soil contaminated with pesticides may also be amenable to remediation. Diuron-fortified moist soil (100 mg/kg) was spread thinly and amended with TiO2 (Higarashi and Jardim 2000). After subjecting the soil to 30 hr of sunlight irradiation, diuron dissipated with a half-life of 10 hr and could no longer be detected after 40 hr. Tordon 101 (a mixture of picloram and 2,4-D) and pentachlorophenol were also susceptible to TiO2-catalyzed photolysis, but they degraded more slowly than diuron. C. Biological Treatment Small waste generators need an easily adaptable, low-cost technology for cleanup of contaminated soils, and biologically based processes should be given the highest priority for development. From a practical standpoint, microbial detoxification systems are cheaper than physical and chemical systems, are easily transportable, not technically complex, and are socially acceptable. Biologically based strategies for disposal and cleanup of contaminated soils include the following (modified from Felsot 1998b): 1. Pretreatment of contaminants with various reagents to produce degradates more amenable to microbial mineralization (i.e., integrated systems; Felsot 1996) 2. Composting alone or in combination with a preliminary adsorption system 3. Treatment of soil with pesticide-degrading microorganisms and/or enzymes (bioaugmentation) 4. Amendment of soil and groundwater with nutrients or alteration of physical conditions to stimulate indigenous microbial metabolism (biostimulation) 5. Use of rhizosphere microbial ecology and root physiology either to metabolize pesticides or to immobilize them by plant uptake with subsequent plant metabolism (phytoremediation) 6. Spreading of waste on agricultural or noncropped land to stimulate degradation, transformation, or immobilization of contaminants (variously known as
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landfarming, land treatment, land application, or land spreading) (Felsot et al. 1995) Integrated physicochemical and biological systems were examined in the context of carbon/lignocellulosic adsorption systems (Section VII, B., 2) and UV/ozonation (Section VII C, 4, ozonation). Among the remaining strategies, composting and landfarming have already been implemented on a commercial scale and seem most amenable to small waste generators. Bioaugmentation and biostimulation techniques are still quite experimental but have been employed by large waste generators at Superfund sites or at manufacturing facilities to effect remediation. Phytoremediation is still in its infancy and has little field testing. Following is an overview of composting, landfarming, phytoremediation, bioaugmentation, and biostimulation from the perspective of laboratory testing or field implementation and the likelihood of implementation for soil cleanup by small waste generators. 1. Composting. Composting may be one solution for remediating pesticidecontaminated soils and organic solids (e.g., sludges, plant foliage). This technology has been applied for generations to the recycling of organic wastes, and relies upon the tremendous increase in microbial activities that occurs when moist organic material is gathered into a large enough mass with sufficient aeration. Naturally occurring microorganisms dramatically increase in numbers and activity as they begin to decompose readily degradable organic substrates. In addition to producing large quantities of carbon dioxide, the resulting metabolic heat often raises the temperature of the compost as high as 60–70 °C. As the temperature of the compost increases, a microbial succession of mesophilic organisms to a thermophilic community occurs. Most investigations of the fate of pesticides and other contaminants during the composting process, which have involved either laboratory- or pilot-scale operations, have focused on the disposition of trace residue levels (1–10 mg/ kg) and potential appearance in finished composts. For example, the insecticides carbaryl, diazinon, and parathion were reported to undergo nearly complete degradation during composting of sewage sludge (Racke 1989), dairy manure (Petruska et al. 1985), and cannery wastes (Rose and Mercer 1968), respectively. The major degradation product for carbaryl and diazinon was reported to be unextractable, bound residues (70%–95%), and little carbon mineralization was noted (<5%). Because of the heavy urban uses of compost for horticultural crops and gardens, much interest has been generated in studies of fate and persistence of turfapplied pesticides in compost (Fogarty and Tuovinen 1991; Buyuksonmez et al. 2000). The complete degradation of chlorpyrifos, isofenphos, diazinon, and pendimethalin was reported to occur within 4 wk after application to turf and 6 wk during composting of the grass clippings (Lemmon and Pylypiw 1992). However, timing of application and subsequent composting start date affected the length of persistence in the compost. For example, late-summer applications
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of the pesticides required 17 wk for complete loss of parent compounds compared to less than 4 wk for early-summer applications. The residual persistence of herbicides in composted grass clippings is of special interest owing to the possible phytotoxicity of the applied compost (Bugbee and Sarceno 1994). After 50 d of composting yard trimmings of leaves and grass amended with radiolabeled 2,4-D (initial concentration, 17 mg/kg), extensive mineralization (47%) was noted, with 43% forming bound residues (Michel et al. 1995). About 1% of the added radiolabel was organosoluble, suggesting some residual parent compound. Other studies have indicated that certain herbicide residues left in finished compost may approach phytotoxic levels. For example, 2,4-D and clopyralid in grass clippings (initial concentration range, 1.5–183 mg/kg) occurred at levels ranging from 0.1 to 1.4 mg/kg in composted grass clippings after 365 d (Vandervoort et al. 1997). Owing to the comparatively high water solublility and thus the likely bioavailability of many of the auxin agonist herbicides, susceptible crops such as tomatoes, grapes, and beans could be injured by use of herbicidecontaining compost as a mulch or as a growth medium even though extensive degradation may have occurred during composting (Bugbee and Saraceno 1994). Promising findings have been reported during composting of sewage sludge contaminated with more recalcitrant polyaromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) (Hogan et al. 1988). However, other reports have demonstrated that some pesticides may persist largely unchanged during composting processes (Mu¨ller and Korte 1974, 1975; Racke 1989; Petruska et al. 1985). For example, little or no degradation of the pesticides aldrin, dieldrin, or monolinuron was noted during composting of sewage sludge waste; limited transformation of buturon and heptachlor (largely to heptachlor epoxide and hydroxychlordene) was noted (Mu¨ller and Korte 1974, 1975). Similarly, chlordane has been found in finished compost from various municipalities in the U.S. that use grass clippings as a raw material (Strom 2000). However, the source of the chlordane is believed to be soil contamination from urban termite treatments because chlordane has not been permitted for use since 1988. Nevertheless, the composting process, which incorporated the contaminated soil, was not capable of effectively degrading the chlordane. Although the cyclodiene insecticides seem recalcitrant to degradation in the specific composting systems studied thus far, further research should explore the role of anerobic decomposition that takes place within full-scale composting systems (Buyuksonmez et al. 2000). Recalcitrant polyhalogenated hydrocarbons have been shown to be metabolized preferentially under anaerobic conditions (Castro and Yoshida 1974; Brown et al. 1987; Adriaens et al. 1995; Zwiernik et al. 1998). Other research has revealed that the reported “disappearance” of some of the more highly persistent pesticides during composting may be largely due to volatilization (Mu¨ller and Korte 1974; Petruska et al. 1985). For example, onehalf of the chlordane lost during sewage sludge composting was captured as
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unchanged, volatilized material (Petruska et al. 1985). Similarly, up to 22% of diazinon in composting cow manure and sawdust was volatilized. In contrast, little volatilization (0.2%) of 10 mg/kg diazinon was reported during composting of deciduous treee leaves and grass clippings (Michel et al. 1997) . Most of the diazinon (36%) was hydrolyzed to 2-isopropyl-6-methyl-4-hydroxypyrimidine (IMHP), and 32% was classified as bound or unextractable. Only recently has composting been evaluated for degrading pesticides and other organic contaminants as a means of waste disposal. Most available reports concern sawmill wastes and soils contaminated with chlorinated phenolic (CP) compounds. Pilot-scale composting (7500 kg and 13 m3) of soil contaminated with 2,4,6-trichlorophenol, 2,3,4,6-tetrachlorophenol, and pentachlorophenol (PCP) at initial CP levels of 437–1108 mg/kg resulted in greater than 90% destruction of these contaminants (Laine and Jorgensen 1997). Parallel benchscale research indicated that considerable mineralization occurred, accounting for on average 60% of the PCP removal. Follow-up efforts with contaminated sawmill soil, which included analyses of PCDDs and PCDFs, confirmed the favorable degradation rate for CP products but revealed that concentrations of PCDDs and PCDFs did not change significantly during composting (Laine et al. 1997). Although earlier, laboratory-scale work conducted at lower concentrations (50 mg/kg) of PCP had indicated that degradation of this CP proceeded at about the same rate during composting as under ambient conditions in sewage sludge, the much higher concentrations in the sawmill waste may have encouraged more extensive microbial adaptation and degradation (Sikora et al. 1982). For high concentrations of recalictrant CP compounds typically present in wood preservative waste, degradation may also be enhanced during composting by additions of adapted microorganisms and fertilizers (Valo and Salkinoja-Salonen 1986). DDT-contaminated soils from old cattle dip sites and PCP-contaminated soils from old timber preservation sites have been composted with green waste in Australia (Singleton et al. 1998). More than 90% removal of DDT (70 mg/kg) occurred during 60 d of thermally heated composting. DDE production was nil, and DDD was the major metabolite. PCP (70 mg/kg) degradation of 90% was achieved during 15 wk of composting without external heating. Seed germination bioassays confirmed the disappearance of PCP-induced toxicity during the composting process. Cattle dip (500,000 L) containing amitraz was remediated in a two stage process using carbon filtration and composting (Van Zwieten et al. 1998). Filtration through the proprietary system removed more than 99% of the amitraz in the cattle dip fluid (148 mg/L). The resulting sludge was composted with sawdust, cow manure, green waste, blood, and bone. Within 104 d, sludge containing 483 mg/kg amitraz was remediated to levels below the detection limit (0.5 mg/kg). Although composting has promise for the destruction of some types of pesticide wastes, additional research is clearly required on the efficiency of this process for a larger variety of pesticides under large-scale, field conditions and
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for compounds at the higher concentrations that characterize pesticide spills and historical rinse-out areas. Despite the need for more research on composting, low-level contamination (<20 mg/kg) of organophosphate and methyl carbamate insecticides and phenoxy and dinitroaniline herbicides may be amenable to composting by small waste generators. The progress of the process could be checked with plant and insect bioassays. When toxicity is nil, the compost could be spread thinly on cropland. 2. Landfarming Landfarming is an effective alternative to composting that could be readily implemented for compounds known to be biotransformed or biomineralized (Felsot 1998b). Soil is excavated and spread thinly on uncontaminated soil, allowing natural chemical and biological processes to transform and degrade the contaminants. Criteria for efficacious landfarming include degradation of pesticides at rates similar to or faster than rates seen for routine field sprays of the same compounds. Also, land receiving the contaminated soil should be managed to avoid excessive runoff and erosion. To meet these criteria, the loading rate of the contaminated soil should be scaled to give concentrations of pesticides no higher than legal application rates. This objective would also protect against phytotoxicity to a crop grown on the land receiving the contaminated soil. Pilot-scale landfarming in different soil types has been tested for creosote and pentachlorophenol wood preserving sludges (Borazjani et al. 1990). Herbicidecontaminated soils arising from a pesticide warehouse fire and a site with frequent dumping of contaminated rinsewater have been landfarmed in field plots planted with soybeans and corn (Felsot et al. 1994, 1995). Landfarming was enhanced by the addition of organic amendments such as cornmeal or sewage sludge (Felsot et al. 1995). Landfarming can be very inexpensive but does require the use of excavation equipment, a tractor, and some type of spreader (e.g., a manure spreader). Encouraging the establishment of a cover crop and thus an active rhizosphere could be beneficial in enhancing degradation (Anderson et al. 1993). Irrigation might be required in very dry climates. An economic analysis in the U.S. showed landfarming under controlled engineering management cost US$20–25 per tonne, whereas incineration cost US$80–90 (Andrews Environmental Engineering, Inc. 1994). When landfarming herbicide-contaminated soils, special care must be taken to not overapply materials to cropland designated for planting of susceptible crops. Despite these limitations, when a contaminated site presents an imminent hazard to a water resource, landfarming is still the quickest and cheapest option for cleanup. An extensive list of best management practices for landfarming pesticide-contaminated soils has been published (Felsot 1998b). 3. Phytoremediation Phytoremediation has been defined as “the use of green plants and their associated microbiota, soil amendments, and agronomic tech-
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niques to remove, contain, or render harmless environmental contaminants” (Cunningham et al. 1996). Phytoremediation is essentially bioremediation with a focus expanded from solely soil microrganisms to plants or plant parts. The word was only coined within the last decade, but as evidenced by a growing number of critical reviews, it has garnered intense interest as an emerging technology for remediation of contaminated soil and groundwater (Anderson et al. 1993; Anderson and Coats 1994; Schnoor et al. 1995; Cunningham et al. 1996; Kruger et al. 1997; Arthur and Coats 1998; Flatham and Lanza 1998; Siciliano and Germida 1998; Gleba et al. 1999; Macek et al. 2000). Conceptually, phytoremediation can be divided into two main processes, phytodecontamination and phytostabilization (Cunningham et al. 1996). Phytodecontamination includes rhizodegradation (rhizosphere-enhanced biodegradation), phytoextraction (i.e., plant uptake), phytodegradation (i.e., plant detoxification), and phytovolatilization (volatilization through leaf surfaces). Phytostabilization includes soil sequestration (i.e., enhanced sorption), lignification (sequestration in cell wall lignins), and humification (enzyme-mediated binding into soil humus). Phytostabilization has received minimal attention for remediation of pesticide contaminants but may be applicable to nonessential metals. Phytodecontamination, especially processes related to rhizosphere-mediated biodegradation, phytoextraction, and phytodegradation have been studied for a handful of pesticides (Table 12). Thus far, pesticide decontamination by phytoremediation has been explored in laboratory studies employing either whole soil (with and without plants), soil suspensions, or water. Other than attempts to phytoextract atrazine from groundwater in a riparian buffer zone (Schnoor et al. 1995), pesticide phytoremediation has not been tested in the field. Based on the laboratory studies, one of the most promising uses of phytoremediation may be cleanup of contaminated ponds or other surface impoundment by aquatic plants (Rice et al. 1997). Even DDT seems subject to uptake and reductive dechlorination by several aquatic plant species (Gao et al. 2000b; Garrison et al. 2000), and research should be expanded to other recalcitrant halogenated contaminants. Despite not knowing the field performance of phytoremediation, crops should still be planted during landfarming operations if levels of waste pesticide residues are kept below phytotoxic levels by appropriate dilution of contaminated soil with field soil (Felsot 1998b). Whether the crops enhance decontamination of the soil or not, at least the small waste generator can recover some of his costs by harvesting a marketable commodity. 4. Bioaugmentation Numerous studies have been published on microbial degradation of a plethora of pesticides. These studies are largely in vitro; enrichment cultures are used to isolate the specific pesticide-degrading microbes or consortia of two or more microbes. In aerobic and anaerobic laboratory cultures, the pesticides are quickly metabolized to lower toxicity transformation products. A good example of this potential under in vitro conditions is the extensive metabolism of DDT, normally considered quite resistant to biodegradation, by
Table 12. Pesticides studied for potential cleanup by phytoremediation. Pesticide (initial concentration)
Phytoremediation process
Plant
Results
Reference
Populus deltoides x nigra (hybrid poplar)
54% sequestration in plant; hydroponic medium
Burken and Schnoor 1998
Atrazine (200 µg/L)
Rhizodegradation Phytoextraction Phytodegradation
Ceratophyllum demersum (hornwort); Elodea canadensis (pondweed); Lemna minor (duckweed)
Half-lives in vegetated pond water, 12–78 d; in nonvegetated water, half-life 144 d; 1%–12% of radiolabel associated with plants after 16 d
Rice et al. 1997
Atrazine (10 mg/kg in mixture with metolachlor and trifluralin)
Rhizodegradation
Kochia sp.
55% remaining in rhizosphere soil compared to 75% in nonrhizosphere soil after 14 d
Anderson et al. 1994
Atrazine (50 mg/kg)
Rhizodegradation
Kochia scoperia (14 other plant species also tested)
62% mineralization in rhizosphere soil in 50 d; 3.5% in sterile control; large differences in mineralization rate among plant species
Anderson and Coats 1995
Atrazine (0.06 mg/kg)
Phytoextraction Rhizodegradation Phytodegradation
Populus deltoides x nigra (hybrid poplar)
92% uptake from silica sand; 20% uptake from silt loam in 22 d; 35% mineralilzation in root exudateamended microcosms after 150 d; 90% of parent atrazine metabolized in leaf tissue
Burken and Schnoor 1996, 1997
Atrazine (50 mg/kg)
Rhizodegradation
Kochia scoparia
Half-life 60 d in rhizosphere soil compared to 193 d in nonrhizosphere
Arthur et al. 2000
soil
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Phytoextraction
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Atrazine (0.27 mg/L)
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Table 12. (Continued.) Pesticide (initial concentration)
Phytoremediation process
Plant
Results
Reference
Phytoextraction
Salix nigra (black willow)
59 mg/d/tree removal rate from irrigated soil
Conger and Portier 1997
Crufomate
Phytoextraction Phytodegradation
Parrot feather (Myriophyllum aquaticum) Duckweed (Spirodela oligorrhiza) Elodea (Elodea canadensis
Half-life for uptake and dissipation of crufomate from solution ranged from 99 to 2302 hr
Gao et al. 2000a
Diazinon (5 mg/kg)
Rhizodegradation
Phaseolus sp.
Mineralization in plant root vs. no root soil-filled tubes significantly enhanced (13% vs. 5%)
Hsu and Bartha 1979
Demeton (10 mg/kg)
Phytoextraction Phytodegradation
Parrot feather (Myriophyllum aquaticum) Duckweed (Spirodela oligorrhiza) Elodea (Elodea canadensis)
Half-life for uptake and dissipation of demeton from solution ranged from 12 to 69 hr
Gao et al. 2000a
DDT (1 mg/L)
Phytoextraction Phytodegradation
Elodea canadensis
Half-life of 3 d for p,p′-DDT in lake sediment extract containing plant; metabolism to DDD (up to 64% in 8 d) and bound residues
Garrison et al. 2000
A.S. Felsot, K. Racke, and D. Hamilton
Bentazon (150 mg/L daily dose for 3 wk through irrigation system)
Table 12. (Continued.) Pesticide (initial concentration)
Phytoremediation process
Plant
Results
Reference
Parrot feather (Myriophyllum aquaticum) Duckweed (Spirodela oligorrhiza) Elodea (Elodea canadensis)
Half-life for uptake of DDT by axenic plant cultures and disappearance from water column ranged from 4.3 to 6.3 hours; in 6 d, no DDT was detected in the water; DDD was a major metabolite
Gao et al. 2000b
Lindane (1 mg/kg)
Rhizodegradation
Oryza sativa (rice)
46% dissipation in rhizosphere soil compared to <5% in nonrhizosphere soil
Brahmaprakash et al. 1985
Malathion (1 mg/kg)
Phytoextraction Phytodegradation
Parrot feather (Myriophyllum aquaticum) Duckweed (Spirodela oligorrhiza) Elodea (Elodea canadensis)
Half-life for uptake and dissipation of malathion from solution ranged from 58 to 116 hr
Gao et al. 2000a
Metolachlor (10 mg/kg in mixture with atrazine and trifluralin)
Rhizodegradation
Kochia sp.
50% remaining in rhizosphere soil compared to 70% in nonrhizosphere soil after 14 d
Anderson et al. 1994
Metolachlor (200 µg/L)
Rhizodegradation Phytoextraction Phytodegradation
Ceratophyllum demersum (hornwort); Elodea (pondweed); Lemna minor (duckweed)
Half-lives in vegetated pond water, 3–8 d; in nonvegetated water, halflife 32 d; 8%–23% of radiolabel associated with plants after 16 d
Rice et al. 1997
Parathion
Rhizodegradation
Phaseolus sp. (bean)
Mineralization significantly enhanced in plant root vs. no root soil-filled tubes (18% vs. 8%)
Hsu and Bartha 1979
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Phytoextraction Phytodegradation
Pesticide Waste
DDT (1 mg/L)
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Table 12. (Continued.) Phytoremediation process
Plant
Results
Reference
Pentachlorophenol (50 mg/L)
Phytoextraction
Populus deltoides x nigra (hybrid poplar)
29% sequestration in plant; hydroponic medium
Burken and Schnoor 1998
Pentachlorophenol (100 mg/kg)
Rhizodegradation Phytoextraction Phytodegradation
Agropyron desertorum (crested wheat grass)
22% mineralized compared to 6% in nonplant system; 35% of radiolabel associated with plant tissue
Ferro et al. 1994
Simazine (0.24 mg/L)
Phytoextraction
Canna hybrida
70% uptake from hydroponic solution in 7 d
Wilson et al. 1999
2,4,5-T (0.66 µg/microcosm)
Rhizodegradation
Sunflower Timothy grass Clover
18% mineralized in planted soil vs. 13% in nonplanted soil
Boyle and Shann 1998
Trifluralin (10 mg/kg in mixture with atrazine and metolachlor)
Rhizodegradation
Kochia sp.
30% remaining in rhizosphere soil compared to 50% in nonrhizosphere soil after 14 d
Anderson et al. 1994
A.S. Felsot, K. Racke, and D. Hamilton
Pesticide (initial concentration)
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an anaerobic microbial culture followed by an aerobic culture (Pfaender and Alexander 1972). Some pesticides can be metabolized by specific bacterial strains as sole carbon or nutrient sources. These types of microbial reactions are called mineralizing because they tend to release large amounts of CO2 during metabolism. These capabilities can be exploited for efficient detoxification of contaminants. Successful use of the variously adapted single-species and consortia microbial cultures to remediate pesticide-contaminated soils under laboratory conditions are less well studied than the demonstration of capability for in vitro degradation. Even rarer are demonstrations of the capability for microbial inoculations (i.e., bioaugmentation) to remediate pesticide-contaminated soils under field conditions. One problem with bioaugmenting contaminated soil with pesticide-degrading microbial inocula is establishment of the microbes in the presence of native microbial populations. Populations of introduced bacteria, for example, may be reduced because of their susceptibiiltiy to predation, starvation, or possibly anitbiotic-producing or lytic microbes (Acea et al. 1988). Another limitation to successful bioaugmentation occurs on aging of soil contaminants. The aging effect has been hypothesized to represent diffusion into soil micropores of pesticide residues that become less accessible to solution phase over time (Pignatello and Huang 1991; Pignatello et al. 1993). Thus, an atrazine-mineralizing bacterial isolate became less effective when inoculated into soil with atrazine residues aged for increasing lengths of time (Radosevich et al. 1997). Where soils are heavily contaminated with pesticides at concentrations that would be toxic to plants desirable for in situ phytoremdiation or stabilization, establishment of pesticide-degrading microbial populations would be beneficial. Successes have been achieved in protecting plants from the phytotoxicity of dicamba and pentachlorophenol by inoculating soil with single-isolate cultures capable of mineralizing the pesticides (Krueger et al. 1991; Pfender 1996). In one case where it was essential to excavate soil contaminated with high levels of dinoseb, an anaerobic treatment process was set up on site in an old train container car. The contaminated soil was mixed with potato starch and inoculated with a dinoseb-adapated microbial culture (Roberts et al. 1993). Metabolism of the potato starch consumed the oxygen in the car, creating anaerobic conditions that were ideal for dinoseb metabolism by the inoculum. Perhaps the most studied microbial inoculation technique for enhancing biodegradation is the use of the white-rot fungi, including Phanerochaete chrysosporium and several other species. White-rot fungi are unique among other microbiota because they can rapidly depolymerize and mineralize lignin, a complex, irrregular nonhydrolyzable wood polymer largely composed of nonrepeating phenylpropane subunits (Hammel 1989). Lignin degadation occurs during secondary fungal metabolism at high rates only during conditions of nitrogen deficiency. The reaction is cometabolic because a primary energy source, for example, cellulose or glucose, is required. The mechanism underlying the ability to rapidly metabolize lignin is believed to be linked to the proven ability
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of the fungus to metabolize a variety of recalcitrant halogenated aromatic compounds and PAHs (Barr and Aust 1994). Capabilities for metabolism have been tested in soil and liquid cultures containing DDT, many chlorinated cyclodiene insecticides, 2,4,5-T, atrazine, PCP, and creosote constituents (i.e., PAHs) (Bumpus et al. 1985; Bumpus and Aust 1987; Kennedy et al. 1990; Mougin et al. 1997; Tuomela et al. 1999). The metabolic mechanisms of the white-rot fungi give the group advantages for soil bioremediation efforts lacking in soil bacteria. The enzymes responsible for lignin degradation, lignin peroxidase and manganese-dependent lignin peroxidase, are extracellularly secreted heme-containing catalysts. Secreting also the oxidant vertaryl alcohol and the reductant oxalate, white-rot fungi degrade lignin and chlorinated contaminants via a unique oxidation–reduction reaction. The degradative reaction occurs analogously to a chemical reaction, that is, no binding of the substrate to the enzyme is required (Barr and Aust 1994), and the degradative reaction is essentially caused by extracellular electron transfers to the contaminant (or lignin). Thus, hydrophobic contaminants of low bioavailability to the intercellular metabolism of bacteria would be susceptible to extracellular biochemical reactions (Barr and Aust 1994). Another advantage of the white-rot fungi is that under aerobic conditions they have been hypothesized to first reduce hydrophobic contaminants such as DDT using a nonlignolytic plasma membrane-dependent redox mechanism. Such initial reduction is necessary to effect oxidative metabolism of many halogenated aromatic hydrocarbons because they are intially in a highly oxidized state. Proof of this mechanism is isolation of DDD in the early stages of DDT degradation in nonlignolytic liquid cultures of P. chrysosporium (Bumpus et al. 1985). When the cultures become nitrogen deficient and start producing lignin peroxidases, the DDD rapidly dissipates, with 13% of the compound mineralized to CO2. Most studies with white-rot fungi have used either in vitro liquid cultures or small amounts of soil in flasks. Two field-scale pilot projects, however, have proven that soils with PCP- and creosote-contaminated wood preservative sludges could be remediated by bioaugmentation with the fungus Phanerochaete sordida. PCP at an initial concentration of 672 mg/kg declined 82% within 60 d after inoculation (Lamar et al. 1993). Concentrations of 3- and 4-ring PAH constituents of creosote declined by a maximum of 95% and 72%, respectively, within 56 d of P. sordida inoculation (Davis et al 1993). In addition to micobial inocula, soils have been amended with cell-free enzymes that degrade specific pesticides. Early studies showed that hydrolases could be added to soils to effect degradation of various organophosphate (OP) insecticides (Munnecke 1976, 1980; Barik and Munnecke 1982). Some success was realized with immobilization of degradative enzymes on inert surfaces through which OP-containing wastewater was passed and detoxified. Spills of diazinon in soil were successfully cleaned up by addition of a hydrolase preparation (Honeycutt et al. 1984). Enzymes that couple phenolic and anilinic com-
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pounds to humic materials can serve to effectively detoxify contaminated soils (Bollag 1992). Minced horseradish roots, which are rich in peroxidase, were used in combination with hydrogen peroxide to remove up to 99% of various phenolic compounds, including PCP, from contaminated water (Roper et al. 1996). One potential problem with direct inoculation of soils or turbid water with enzymes is possible inactivation or sorption of the enzyme to clay, making it nonfunctional (Masaphy et al. 1996). Although early reviews showed promise of cell-free microbial enzymes for degrading in situ organophosphate, carbamate, phenylurea, acylanilide, and phenoxyacetate pesticides (Johnson and Talbot 1983), little field evidence has been collected to show the method feasible for field-scale remediation. To circumvent the problem of successful establishment of microbial inocula in native soil, soils could be washed and the effluent transferred to a slurryphase biofilm reactor. Slurry phases are essentially amorphous mixtures of contaminated sediment or soil mixed with adapted microbial inocula or sufficient nutrients and kept under optimal temperature conditions in a confined reactor cell. Biofilms contain fixed particles in a reactor with contaminated water and nutrients flowing around and across them. A microbial consortia develops on the enhanced surface area of the particles. Significant biodegradation of creosote- and PCP-contaminated sediment and soil has been observed in a slurryphase bioreactor (Mueller et al. 1991). PCP- and chlordane-contaminated water were remediated in a biofilm reactor (Sabatini et al. 1990; Karamanev and Samson 1998). In a novel innovation, microbial-inhibiting concentrations of PCP (300 mg/L) were efficiently degraded (>95%) within 6 d in a bioreactor consisting of polyurethane-immobilized Flavobacterium cells (O’Reilly and Crawford 1989). The polyurethane presumably sorbs the PCP, effectively lowering its solution concentration below toxic levels. Success at enhancing biodegradation has also been achieved by inoculating soils with small quantities of previously exposed soil. For example, atrazine mineralization was increased from 2%–3% to 25%–30% in a constructed wetland sediment inoculated to a level of 1% (w/w) with soil previously exposed to atrazine spills (Runes et al. 2001b). Simple organic amendments (cellulose and cattail leaves) failed to enhance atrazine mineralization. Bioaugmentation with this “acclimated” soil also enhanced atrazine mineralization under flooded conditions. In summary, pesticide-degrading microorganisms and cell-free enzymes have been added to soil in an attempt to enhance the degradation of several groups of pesticides. Much of the work has been conducted under laboratory conditions. With some exceptions, such as, the use of white-rot fungi to clean up field soil contaminated with wood preservative waste and the use of hydrolase to clean up diazinon-contaminated soil, not enough development on a pilot scale has been conducted to make bioaugmentation a usable technique for small waste generators.
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5. Biostimulation Most pesticides are detoxified in soil by cometabolic reactions. Cometabolized compounds are partially transformed during secondary metabolism, but a primary carbon or nutrient source is required to sustain the microbial cells (Alexander 1985). The transformation products are usually much less acutely toxic than the parent compound. A significant proportion of the transformation products are either susceptible to further microbial metabolism and/or covalently incorporated into soil organic matter as a bound residue with a consequent loss of bioavailability. Thus, one strategy for decontaminating soils is to enhance native microbial metabolism by stimulating population growth of microorganisms that can transform the pesticide. Degradative activities of resident microflora can be stimulated by enriching the soil environment with nutrient amendments or by changing the physical characteristics of the environment. For example, organic and mineral amendments enhanced the degradation of a variety of pesticide classes in the sediment of a pit that had been used to store rinsewater from spray application equipment (Winterlin et al. 1989). Chlorinated pesticides such as toxaphene were partially decontaminated by addition of cornmeal or cotton gin waste to soils maintained under anaerobic conditions (Mirsatari et al. 1987). DDT and lindane degraded significantly faster in flooded soil amended with rice straw or other organic matter than in unamended soil (Mitra and Raghu 1986; Castro and Yoshida 1974). Degradation of PCP in anaerobic (flooded) soil was enhanced by the addition of anaerobic sewage sludge (Mikesell and Boyd 1988). Similarly, beta hexachlorocyclohexane, a significant noninsecticidal, biologically recalcitrant contaminant in technical lindane formulations, was degraded when soil was amended with an anaerobic sludge and kept under methanogenic conditions (Van Eekert et al. 1998). In each case, the sludge not only added a usable source of carbon but also bioaugmented the soil by transferring anerobic bacteria capable of dechlorinating the pesticides. Despite apparent successes in enhancing biodegradation through biostimulation, the technique may not be universally applicable. For example, alachlor at levels of 100 mg/kg dissipated significantly faster in soil amended with cornmeal or soybean meal than in unamended soil (Felsot and Dzantor 1990). However, at a contamination level of 1000 mg/kg, which would not be unusual for a surface spill of formulated herbicide, dissipation was not enhanced by organic amendments. Furthermore, the effectiveness of biostimulation was reduced when cornmeal was applied to contaminated soil with aged alachlor residues compared with freshly added residues (Felsot and Dzantor 1997). Although the composting process itself may enhance the degradation of some types of pesticides, addition of finished compost to soil inhibited mineralization of several triazine herbicides freshly added to soil at levels below1 mg/kg (Barriuso et al. 1997). On the other hand, compost added to soil from a contaminated field site enhanced the degradation of aged residues of trifluralin, metolachlor, and pendimethalin and reduced phytoxicity of the herbicides (Cole et al. 1995). Under aerobic conditions, glucose and wheat straw have enhanced the biodegra-
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dation of the herbicide metribuzin, but alfalfa residue slowed biodegradation (Pettygrove and Naylor 1985). Thus, the initial concentration of the pesticide contaminants and the composition of the organic amendments significantly influence the efficiency of biostimulation as a waste remediation tool. Despite the potential limitations of biostimulation, it seems to be a low-cost technology using readily available materials (e.g., cornmeal, straw) that can be mixed into the soil to increase microbial activity and coincidentally effect biotransformations leading to detoxification. Biostimulation has been used with some success in conjunction with landfarming; combining the techniques may overcome some of the problems with very high concentrations that could be toxic to native microflora. However, further research is needed to elucidate amendment materials that might slow down pesticide degradation.
IX. Conclusions Pesticide waste disposal and decontamination is made very difficult by the presence of a diversity of chemistries at extraordinarily high concentrations. Research on pesticide waste disposal has focused mainly on laboratory-scale studies, often evolving from basic studies on factors influencing pesticide behavior in soil, water, and plants. Of all the technologies reviewed, incineration of obsolete stocks has been put into practice, especially to help developing countries clean up highly contaminated sites. Because of historically deficient wastewater handling practices, pesticide-contaminated soils occur throughout the world. Yet, for the small waste generator, all the potentially useful chemical and biological cleanup methods have not yet been adequately field tested or field proven to make them reliable, easy, and affordable to implement. Considering that a wide diversity of pesticides will long continue to be necessary for crop protection, small waste generators must adhere earnestly to proposed guidelines for waste management. The guidelines can be ranked from least to most preferred options for waste management. Given the problems in disposing and treating waste, especially after soil has become contaminated, practices to avoid or minimize waste and recycle products and rinsewater must be the top objectives. Yet, spills will inevitably occur, and soils will become contaminated. In that event, quick cleanup using landfarming methods is most likely to be affordable and easily implemented by small waste generators. To maximize results, waste generators can use a combination of biostimulation and phytoremediation to speed up contaminant detoxification. Meanwhile, research should continue with bioremediation techniques and greater efforts should be given to field testing of innovations once they have been proven in the lab.
X. Consensus Recommendations of the IUPAC Commission on Agrochemicals and the Environment A. General Waste Management Strategy 1. All potential pesticide waste generators should develop a total management strategy for waste that is applicable to their specific operation. Such a strategy would essentially follow the waste management hierarchy proposed by GCPF
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(formerly GIFAP) or the Australian Code of Practice for Agriculture Waste Management. The essential elements of this hierarchy, in order of most desired to least preferred course of action, include waste avoidance, waste reduction, waste recycling, waste treatment, and waste disposal. B. Obsolete Pesticide Stocks 1. Pesticide users should avoid generation of obsolete stocks by following FAO guidelines for inventorying, storing, and managing stocks and by ensuring a regular turnover of products. 2. Pesticide users needing to dispose of obsolete stocks in countries without facilities for disposing of hazardous waste should seek guidance and assistance from their appropriate governmental bodies, which should be familiar with FAO guidelines and the potential to request UN assistance. 3. Obsolete stocks must be secured before disposal and waste components identified. Because concentrations of active ingredients in obsolete stocks are very high, TLC will be an appropriate analytical method for accomplishing tentative identification when product labels are corroded or missing. For some disposal options, an elemental analysis may be required. 4. Currently, the most practical method for disposing of large quantities of obsolete stocks is incineration in a qualified hazardous waste incinerator. Chemical methods of destruction are at best transformations, including dechlorination. More research is needed on alternative chemical destruction methods that could supplant the need for incineration. C. Small Waste Generator Practices for Containers and Rinsewater Disposal 1. For individual pesticide users, all containers should be triple rinsed with the contents poured into the spray tank. Nonrecyclable containers such as bags should be gradually phased out. Open burning of bags should be avoided owing to the generation of potentially bioactive incomplete combustion products. 2. After application, individual pesticide users should dispose of tank rinsates by spraying around edges of their fields. 3. If it is not practical to dispose of rinsates by spraying around the edges of a field, as might be the case with commercial application services, then rinsewater should be saved and used as a diluent for the next spray application. Thus, an organized management plan should be developed for optimizing spray rinsewater storage and recycling. 4. To avoid creation of contaminated soils by spills and dripping during pesticide mixing and loading, pesticide users should consider constructing biobeds over which to carry out product preparation. More research should be conducted to determine the feasibility of using biobeds for disposing of wastewater that cannot be recycled.
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5. Commercial application services that cannot recycle rinsate water should consider separation of active ingredients by a sorbent system such as activated carbon. However, plans must be made for ultimate disposal of the sorbent. At present, incineration seems the only practical option for sorbent disposal, although a limited amount of research has focused on composting. D. Cleanup of Contaminated Soils 1. Soils contaminated with currently registered products can be disposed of by landfarming so long as soil loading rates do not exceed maximum permissible application rates for the pesticidal active ingredients. Regulatory authorities need to clarify waste specification and permitting rules to facilitate landfarming. 2. Composting has achieved some success for remediating pentachlorphenolcontaminated soils, but it is not generally applicable for other types of pesticides until further research defines the ultimate fate and bioavailability of parent chemical and degradation products. 3. Further research should be conducted on phytoremediation of contaminated soils to prove the efficacy of in situ treatment. A greater variety of active ingredients than has been tested to date should be studied, and conditions that optimize degradation need further elucidation. E. Future Research Needs 1. At present, although chemical degradation is a vigorous area of research, especially regarding oxidation processes (UV/ozonation, dark and photoassisted Fenton’s reagent, and photocatalysis with titanium dioxide), it is not applicable for disposal of wastes by individual pesticide users. Chemical degradation techniques need to be developed on a pilot-scale basis and engineered for practical use by the individual user, or at least made available as a mobile unit that can be shared among cooperating users. 2. Chemical degradation techniques may be feasible for pesticide users in the future, especially if research is focused on development of flowthrough systems. 3. Further research is needed to characterize degradation products of chemical treatment to aid in determining the efficiency of methods for final disposition. 4. Biological treatment methods that rely on bioaugmentation and biostimulation are not generally feasible at present for routine disposal of wastewater or remediation of contaminated soil. Further research is needed to improve the efficiency of biological methods and increase the diversity of agrochemicals that may be amenable to biological treatment. 5. Further research should be conducted to integrate chemical and biological treatment methods.
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Summary Generation of pesticide waste is inevitable during every agricultural operation from storage to use and equipment cleanup. Large-scale pesticide manufacturers can afford sophisticated recovery, treatment, and cleanup techniques. Smallscale pesticide users, for example, single farms or small application businesses, struggle with both past waste problems, including contaminated soils, and disposal of unused product and equipment rinsewater. Many of these problems have arisen as a result of inability to properly handle spills during equipment loading and rinsewater generated after application. Small-scale facilities also face continued problems of wastewater handling. Old, obsolete pesticide stocks are a vexing problem in numerous developing countries. Pesticide waste is characterized by high concentrations of a diversity of chemicals and associated adjuvants. Dissipation of chemicals at elevated concentrations is much slower than at lower concentrations, in part because of microbial toxicity and mass transfer limitations. High concentrations of pesticides may also move faster to lower soil depths, especially when pore water becomes saturated with a compound. Thus, if pesticide waste is not properly disposed of, groundwater and surface water contamination become probable. The Waste Management Hierarchy developed as an Australian Code of Practice can serve as a guide for development of a sound waste management plan. In order of desirability, the course of actions include waste avoidance, waste reduction, waste recycling, waste treatment, and waste disposal. Proper management of pesticide stocks, including adequate storage conditions, good inventory practices, and regular turnover of products, will contribute to waste avoidance and reduction over the long term. Farmers can also choose to use registered materials that have the lowest recommended application rates or are applied in the least volume of water. Wastewater that is generated during equipment rinsing can be recycled by spraying it onto cropland, thus avoiding a soil contamination problem. If it is not feasible to spray out rinsates, then water treatment becomes necessary. However, for small waste generators, practical technology is still too experimental and not easily implemented on an individual farm or at a small application business. Nevertheless, research has been quite active in application of advanced oxidation processes (UV/ozonation; photoassisted Fenton reaction; photocatalysis using TiO2). Obsolete pesticide stocks in developing countries are being packaged and shipped to developed countries for incineration. Contaminated soil can also be incinerated, but this is not practical nor affordable for small waste generators. Chemical degradation of chlorinated hydrocarbon pesticides may be amenable to dechlorination by alkali polyethylene glycol treatment, but further study is needed to make the technique practical for small waste generators. Contaminated soils may be amenable to cleanup by one of several biological treatment methods, including composting, landfarming, and bioaugmentation/ biostimulation. Composting and landfarming (which may be used in combina-
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Table 13. Chemical Abstracts Service (CAS) numbers and common and chemical names of pesticides cited in this review. Common name
CAS no.
2,4,5-T 2,4-D Alachlor
93-76-5 94-75-7 15972-60-8
Aldicarb
116-06-3
Aldrin
309-00-2
Amitraz Asulam Atrazine Azinphos-methyl
33089-61-1 3337-71-1 1912-24-9 86-50-0
Benomyl
17804-35-2
Bentazon
25057-89-0
Bromacil Bromophos
314-40-9 2104-96-3
Captan
133-06-2
Carbaryl Carbofuran
63-25-2 1563-66-2
Chloridazon Chlorpyrifos
1698-60-8 2921-88-2
Chlorpyrifosmethyl Coumaphos
5598-13-0
Crufomate
299-86-5
Cyanazine
21725-46-2
DBCP DDT Demeton
96-12-8 50-29-3 8065-48-3
Diazinon
333-41-5
Dicamba
1918-00-9
56-72-4
Chemical name 2,4,5-Trichlorophenoxyacetic acid 2,4-Dichlorophenoxyacetic acid 2-Chloro-2′,6′-diethyl-N(methoxymethyl)acetanilide 2-Methyl-2-(methylthio)propionaldehyde O-methylcarbamoyloxime 1,2,3,4,10,10-Hexachloro-1,4,4a,5,8,8a-hexahydroendo,exo-1,4:5,8-dimethanonaphthalene N,N′-[(Methylimino) dimethylidyne]di-2,4-xylidine 4-Amino-benzosulfonyl-methylcarbamate 2-Chloro-4-ethylamino-6-isopropylamino-S-triazine S-(3,4-Dihydro-4-oxobenzo[d]-[1,2,3]-triazin-3ylmethyl) O,O-dimethyl phosphordithioate Methyl-1-[(Butylamino)carbonyl]-H-benzimidazol2-ylcarbamate 3-Isopropyl-1H-2,1,3-benzothiadiazin-4(3H)-one 2,2-dioxide 5-Bromo-3-sec-butyl-6-methyluracil O-(4-Bromo-2,5-dichlorophenyl O,O-dimethyl phosphorothioate 3a,4,7,7a-Tetrahydro-2-[(trichloromethyl)thio]-1Hisoindole-1,3(2H)-dione 1-Naphthyl-N-methylcarbamate 2,3-Dihydro-2,2-dimethylbenzofuran-7-yl methylcarbamate 5-Amino-4-chloro-2-phenylpyridazin-3(2H)-one O,O-Diethyl-O-3,5,6-trichloro-2-pyridyl phosphorothioate O, O-Dimethyl O-(3,5,6-trichloro-2-pyridyl) phosphorothioate 3-Chloro-7-diethoxyphosphinothioyloxy-4methylcoumarin 4-tert-Butyl-2-chlorophenyl methyl methylphosphoramidate 2-(4-Chloro-6-ethylamino-1,3,5-triazin-2-ylamino)2-methylpropionitrile 1,2-Dibromo-3-chloropropane 1,1,1-Trichloro-2,2-bis(4-chlorophenyl) ethane O,O-Diethyl O-[2-(ethylthio)ethyl]phosphorothioate, mixture with O,O-diethyl S-[2-(ethylthio)ethyl]phosphorothioate O,O-Diethyl-O-(2-isopropyl-4-methyl-6pyrimidinyl)- phosphorothioate 3,6-Dichloro-2-methoxybenzoic acid
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Table 13. (Continued). Common name
CAS no.
Dichlofenthion Dieldrin
97-17-6 60-57-1
Dimethoate
60-51-5
Dinoseb Disulfoton
88-85-7 298-04-4
Diuron DSMA EDB Endosulfan
330-54-1 144-21-8 106-93-4 115-29-7
EPN EPTC Fenitrothion
2104-64-5 759-94-4 122-14-5
Fenthion
55-38-9
Imazaquin
81335-37-7
Imidacloprid
13826-41-3
Lindane Linuron Malathion
58-89-9 330-55-2 121-75-5
MCPA Metamitron
94-74-6 41394-05-2
Methamidophos Methidathion
10265-92-6 950-37-8
Methoxychlor Methyl parathion Metobromuron Metolachlor
72-43-5 298-00-0 3060-89-7 51218-45-2
Metribuzin
21087-64-9
Mevinphos
7786-34-7
Paraquat Parathion
4685-14-7 56-38-2
Chemical name O-2,4-Dichlorophenyl O,O-diethyl phosphorothioate 1,2,3,4,10,10-Hexachloro-6,7-epoxy1,4,4a,5,6,7,8,8a-octahydro-endo,exo-4:5, 8-dimethanonaphthalene O,O-Dimethyl S-(N-methylcarbamoylmethyl) phosphorodithioate 4,6-Dinitro-2-sec-butylphenol O,O-Diethyl S-2-(ethylthio)-ethyl phosphorodithioate 3-(3,4-Dichlorophenyl)-1,1-dimethylurea Disodium methanearsonate 1,2-Dibromoethane 6,7,8,9,10,10-Hexachloro-1,5,5a,6,9,9a-hexahydro6,9-methano-2,4,3-benzadioxathiepin-3-oxide O-Ethyl-O-4-nitrophenylphenylphosphonothioate S-Ethyl dipropylthiocarbamate O,O-Dimethyl O-(3-methyl-4-nitrophenyl) phosphorothioate O,O-dimethyl O-4-methylthio-m-tolyl phosphorothiaote 2-[4,5-Dihydro-4-methyl-4-(1-methylethyl)-5-oxo1H-imidazol-2-y1]-3-quinolinecarboxylic acid 1-[(6-Chloro-3-pyridinyl)methyl]-N-nitro-2imidazolidinimine 1,2,3,4,5,6-Hexachlorocyclohexane gamma isomer 3-(3,4-Dichlorophenyl)-1-methoxy-1-methylurea S-[1,2-bis(Ethoxycarbonyl)ethyl]O,O-dimethyl phosphorodithioate 2-Methyl-4-chlorophenoxyacetic acid 4-Amino-4,4-dihydro-3-methyl-6-phenyl-1,2,4triazine-5-one O,S-Dimethylphosphoramidothiolate S-2,3-Dihydro-5-methoxy-2-oxo-1,3,4-thiadiazol-3ylmethyl O,O-dimethylphosphorodithioate 1,1,1-Trichloro-2,2-bis(4-methoxyphenyl)ethane O,O-Dimethyl-O-(4-nitrophenyl) phosphorothioate N′-(4-Bromophenyl)-N-methoxy-N-methylurea 2-Chloro-N-(2-ethyl-6-methylphenyl)-N-(2methoxy-1-methylethyl) acetamide 4-Amino-6-tert-butyl-4,5-dihydro-3-methyltio-1,2,4triazin-5-one 2-Methoxycarbonyl-1-methylvinyl dimethyl phosphate 1,1-Dimethyl-4,4’-bipyridinium ion O-O-Diethyl O-(p-nitrophenyl) phosphorothioate
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Table 13. (Continued). Common name
CAS no.
Pentachlorophenol Phorate Picloram PMA Prometryne
87-86-5 298-02-2 1918-02-1 62-38-4 7287-19-6
Propazine
139-40-2
Propoxur Ronnel
114-26-1 299-84-3
Simazine Thiometon
122-34-9 640-15-3
Trifluralin Zineb Zinophos
1582-09-8 12122-67-7 297972
Chemical name
O,O-Diethyl[(ethylthio)methyl] phosphorodithioate 4-Amino-3,5,6-trichloro-2-picolinic acid Phenylmercuric acetate N2,N4-di-Isopropyl-6-methylthio-1,3,5-triazine-2,4diyldiamine 6-Chloro-N2,N4-di-isporopyl-1,3,5-triazine-2,4diyldiamine 2-Isopropoxyphenyl N-methylcarbamate O,O-Dimethyl O-(2,4,5-trichlorophenyl)phosphorothioate 6-Chloro-N2,N4-diethyl-1,3,5-triazine-2,4-diamine S-[2-(Ethylthio)ethyl] O,O-dimethyl phosphorodithioate α,α,α-Trifluoro-2,6-dinitro-N,N-dipropyl-p-toluidine (Ethylenebis(dithiocarbamate)) zinc O,O-Diethyl-O-(2-pyrazinyl)phosphorothioate
tion with biostimulation) may be the most practical of the biological methods that is immediately ready for implementation by small-scale pesticide waste generators.
Acknowledgments Preparation of this review was supported by funding for Project 640/31/91, Disposal and Degradation of Pesticide Waste, from the Division of Chemistry and the Environment of the International Union of Pure and Applied Chemistry. Members of the Commission for Agrochemicals and the Environment who reviewed the project report and contributed to the consensus recommendations included P.T. Holland, Y.-H. Kim, H.A. Kuiper, B.H.J. Linders, R.D. Wauchope, C. Bellin, R. Dieterle, R.H. Gonzalez, C. Harris, A. Katayama, B. Petersen, S. Reynolds, B. Rubin, G.R. Stephenson, K. Tanaka, J.B. Unsworth, S.-S. Wong, and S.M. Yeh.
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Index
Adsorption coefficient (Kd), methyl bromide, 68 Advisory Committee on Crop Protection Chemistry, 125 Aerodynamic method, methyl bromide volatilization, 85 Air sampling adsorbents, methyl bromide, 56 Air sampling, methyl bromide, 54, 55 Alachlor, degradation by photo-Fenton process, 160 Animal nutrition, essential elements defined, 2 Anthropogenic processes, heavy metals in soils, 4 Apatite, adsorption of heavy metals, 21 Arsenic, soil contaminant sources, 3 Bacteria, methyl bromide biodegradation, 61 Bioaugmentation, pesticide contaminated soils, 170 Biobeds, disposal pesticide spray spills, 149 Biostimulation, pesticide contaminated soils, 178 Biosuper effect, sulfur microbial oxidation effect soil, 12 Bromine residues, above-ground plant tissues, methyl bromide, 66 Bromine residues, edible plant parts, methyl bromide, 67 Bromine residues, plant seedlings, methyl bromide, 66 Bromine residues, soils/plants from methyl bromide fumigation, 62, 64 Bromine vs chlorine, ozone breakdown efficiency, 47 Bromomethane, see methyl bromide, 46 Cadmium, bioaccumulation field crops, 2 Cadmium, concentration in fertilized/unfertilized soils Australia, 19
Cadmium, soil accumulation from fertilizers, 14 Cadmium, soil contaminant sources, 3 Calcium dihydrogen phosphate, fertilizer, 11 Calcium monohydrogen phosphate, fertilizer, 11 Carbofuran, degradation by photo-Fenton process, 160 Carbonate apatite, phosphate fertilizer, 11 CAS numbers, pesticide list, 183 Chamber methods, methyl bromide volatilization, 83 Charcoal tubes, methyl bromide adsorbent, 57 Chloropicrin, added warning agent in methyl bromide, 49 Chloropicrin, methyl bromide replacement, 107 Chromium, soil contaminant sources, 3 Cleanup of pesticide contaminated soils, 161 Commission on Agrochemicals and the Environment, 125 Complexation affinity, metal cations by organic matter (diagram), 7 Composting, pesticide contaminated soils, 166 Container and packaging (pesticide) disposal, 142 Copper, soil contaminant sources, 3 Dehalogenation, pesticide waste disposal, 151 Dichloropropene (1,3-D), methyl bromide replacement, 107 Disposal of unused pesticide stocks, 134 Economic impact, methyl bromide phaseout, 47 Egyptian phosphate rock, fertilizer, 20 Environmental fate, methyl bromide fumigant, 45 ff.
201
202
Index
Environmental fate processes, methyl bromide, 58 Essential elements, plant/animal nutrition, defined, 2 Field experiments, methyl bromide volatilization, 89 Fluroapatite, phosphate fertilizer, 11 Food chain, metals accumulation from waste disposal on soils, 2 Fumigants, mass transfer coefficients across plastic films, 96, 101 Gafsa phosphate rock, fertilizer, 20 Gas chromatography detectors, methyl bromide, 55 GIFAP, (International Group of National Assoc of Mnftrs of Agrochemical Prods), 131 Half-life, methyl bromide, 59, 60, 61 Half-life, methyl bromide in soil, 104 Half-lives, pesticides in soil, 127 Heavy metal adsorption, phosphate-induced, 22 Heavy metal availability in soil-plant system, phosphorus role, 1 ff. Heavy metal immobilization in soils, phosphorus role, 1 ff. Heavy metal mobilization in soils, phosphorus role, 1 ff. Heavy metal precipitation in soils, phosphate-induced, 24 Heavy metal precipitation, phosphate liming action, 27 Heavy metals, adsorption/desorption onto phosphates, 21 Heavy metals, adsorption in soils, 6 Heavy metals, complexation in soils, 7 Heavy metals, defined, 1 Heavy metals, fractionation categories in soils, 8 Heavy metals in soils, anthropogenic processes, 4 Heavy metals in soils, pedogenic processes, 4 Heavy metals in soils, sources (table), 3 Heavy metals, precipitation in soils, 8
Heavy metals, reactions in soils (diagram), 5 Heavy metals, soil immobilization by phosphate compounds, 12, 13 Heavy metals, soil immobilization by water-soluble/insoluble phosphates, 16 Heavy metals, soil mobilization by phosphate compounds, 12, 13 Heavy metals, soil mobilization by watersoluble/insoluble phosphates, 14 Heavy metals, solid-phase speciation, 8 Heavy metals, sources in soil environment, 4 Henry’s law constant, methyl bromide, 68, 69 Henry’s law constants, soil fumigants, 107 Horizontal flux method, methyl bromide volatilization, 88 Human toxicity, methyl bromide, 50 Hydrolysis, methyl bromide in water, 59 Hydrolysis, pesticide waste disposal, 150 Hydroxy apatite, phosphate fertilizer, 11 Incineration, pesticide contaminated soils, 163 Incineration, pesticide disposal method, 136 Incineration, pesticide plastic & paper packaging, 143 International Group of National Assoc of Mnftrs of Agrochemical Prods (GIFAP), 131 International Union of Pure and Applied Chemistry, (IUPAC), 125 IUPAC (International Union of Pure and Applied Chemistry), 125 Kd (adsorption coefficient), methyl bromide, 68 KPEG, see potassium polyethylene glycol ether, 140 Landfarming, pesticide contaminated soil cleanup, 169 Lead, soil contaminant sources, 3 Lead-phosphate interactions, soils, 26 Leaded gasoline, atmospheric methyl bromide source, 53
Index Liming action of phosphate, heavy metal precipitation, 27 Liming, increases metal retention in soils, 8 Manganese, soil contaminant sources, 3 MeBr (methyl bromide), 46 Mercury, soil contaminant sources, 3 Metal cations, complexation affinity by organic matter (diagram), 7 Metal phosphate compounds, water solubility, 25 Metal phosphate precipitation, primary P-induced immobilization, 8 Metal phosphates, soil precipitates, 24 Metal-organic complex formation, factors affecting, 7 Metalloids, defined, 2 Metam sodium, methyl bromide replacement, 107 Methyl bromide (MeBr), 46 Methyl bromide, adsorption coefficient, 68 Methyl bromide, air sampling & analysis, 54, 55 Methyl bromide, air sampling adsorbents, 56 Methyl bromide, air sampling containers, 56 Methyl bromide, application depth effect, 100 Methyl bromide, atmospheric oceanic emissions, 53 Methyl bromide, bacterial biodegradation, 61 Methyl bromide, chemical alternatives, 47 Methyl bromide, chloropicrin warning agent, 49 Methyl bromide, containment to minimize volatilization, 94 Methyl bromide, degradation rate effect on emissions, 104 Methyl bromide, degraded to bromine ion in soil, 82 Methyl bromide, developing alternatives, 106 Methyl bromide, diffusion in air, 78 Methyl bromide, diffusion in soils, 77
203
Methyl bromide, effective soil diffusion coefficient, 75 Methyl bromide, effects of organic matter on hydrolysis, 60 Methyl bromide, elimination economic impact, 47 Methyl bromide, enhancing soil transformation rates, 62 Methyl bromide, environmental concerns, 46 Methyl bromide, environmental fate, 45 ff. Methyl bromide, environmental transformation, 58 Methyl bromide, estimating total loss from soil, 81 Methyl bromide, fumigation application methods, 51 Methyl bromide, fumigation Br residues in soils/plants, 62, 64 Methyl bromide, fumigation soil bulk density effect, 103 Methyl bromide, fumigation soil water content effect, 102 Methyl bromide, fumigation tarp coverings, 51 Methyl bromide, fumigation uses, 48, 50 Methyl bromide, gas chromatography detectors, 55 Methyl bromide, global usage (budget), 47 Methyl bromide, human toxicity, 50 Methyl bromide, hydrolysis in soil, 60 Methyl bromide, hydrolysis water effects, 59 Methyl bromide, injection depth & use of plastic film effects, 99, 101 Methyl bromide, known atmospheric sources, 53 Methyl bromide, methods for minimizing volatilization, 94 Methyl bromide, mobility indices, 73 Methyl bromide, natural atmospheric removal, 53 Methyl bromide, ozone depletion potential, 46, 52 Methyl bromide, phase partitioning, 68 Methyl bromide phase-out, economic concerns, 47
204
Index
Methyl bromide phase-out, major crop loss estimates, 48 Methyl bromide phase-out, Montreal Protocol agreement, 50 Methyl bromide, photohydrolysis, 60 Methyl bromide, physical/chemical properties, 49 Methyl bromide, phytotoxicity, 66 Methyl bromide, plastic film barriers, 95 Methyl bromide, production by growing plants, 67 Methyl bromide, reactions with ozone, 52 Methyl bromide, replacement fumigants, 107 Methyl bromide, simulated environmental fate, 70 Methyl bromide, soil fumigant history, 50 Methyl bromide, soil gas distribution after injection, 78 Methyl bromide, soil gas partitioning, 79 Methyl bromide, soil half-life, 104 Methyl bromide, solubility vs temperature, 69 Methyl bromide, transport in dry soils, 74 Methyl bromide, transport model, 70 Methyl bromide, vapor pressure vs temperature, 69 Methyl bromide, volatilization aerodynamic method, 85 Methyl bromide, volatilization boundary condition, 71 Methyl bromide, volatilization chamber methods, 83 Methyl bromide, volatilization experiments & mass balance, 90 Methyl bromide, volatilization field experiments, 89 Methyl bromide, volatilization from soil, 81 Methyl bromide, volatilization horizontal flux method, 88 Methyl bromide, volatilization micrometeorological methods, 85 Methyl bromide, volatilization theoretical profile shape method, 86 Methyl bromide, volatilization trajectory simulation, 87 Methyl isothiocyanate (MITC), methyl bromide replacement, 107
Methylococcus sp., methyl bromide biodegradation, 61 Methylomonas sp., methyl bromide biodegradation, 61 Micrometeorological methods, methyl bromide volatilization, 85 Mobility indices, methyl bromide, 73 Molybdenum, soil contaminant sources, 3 Montreal Protocol, methyl bromide phaseout agreement, 50 Mycorrihizal fungi colonization of plant roots, rhizosphere microflora, 33 Mycorrhizal rhizosphere, effects on metal tolerance in plants, 31 National Registration Authority, Australia, 129 Nickel, soil contaminant sources, 3 Nitrifying bacteria, methyl bromide biodegradation, 61 Nitrosococcus sp., methyl bromide biodegradation, 61 Nitrosolobus sp., methyl bromide biodegradation, 61 Nitrosomonas sp., methyl bromide biodegradation, 61 Nonessential elements, plant/animal nutrition defined, 2 Nucleophiles, methyl bromide hydrolysis, 59 Obsolete pesticide stocks, IUPAC disposal recommendations, 180 Obsolete pesticides, bulk disposal methods, 135 Organic matter, effects on methyl bromide hydrolysis, 60 Oxidation, pesticide waste disposal, 152 Ozonation, pesticide waste disposal, 152, 154 Ozone Assessment Synthesis Panel (U.N.), 46 Ozone breakdown efficiency, bromine vs chlorine, 47 Ozone depletion, chlorinated compounds, 46 Ozone depletion, methyl bromide, 52 Ozone depletion potential index, 46
Index Ozone depletion potential, methyl bromide, 46 PCBs, degraded by KPEG, 141 PCDDs, produced by chlorinated pesticide incineration, 140 PCDFs, produced by chlorinated pesticide incineration, 140 Pedogenic processes, heavy metals in soils, 4 Pesticide 50% dissipation times (DT50%), in soil Pesticide candidates for phytoremediation cleanup, 171 Pesticide CAS numbers list, 183 Pesticide chemical names list, 183 Pesticide Chemical Abstracts Service (CAS) numbers list 183 Pesticide common names list, 183 Pesticide concentrations in waste, user sites, 126 Pesticide container & rinsewater disposal, IUPAC recommendations, 180 Pesticide container and packaging disposal, 142 Pesticide container recycling, 142 Pesticide contaminated soils, bioaugmentation, 170 Pesticide contaminated soils, biological treatment, 165 Pesticide contaminated soils, biostimulation, 178 Pesticide contaminated soils, cleanup methods, 161 Pesticide contaminated soils, composting, 166 Pesticide contaminated soils, incineration, 163 Pesticide contaminated soils, IUPAC cleanup recommendations, 180 Pesticide contaminated soils, KPEG treatment, 164 Pesticide contaminated soils, landfarming, 169 Pesticide contaminated soils, phytoremediation, 169 Pesticide contaminated soils, soil separation techniques, 162
205
Pesticide contaminated soils, zero-valent iron treatment, 164 Pesticide disposal, 123 ff. Pesticide disposal, bulk quantity FAO guidelines, 135 Pesticide disposal, developing countries, FAO guidelines, 135 Pesticide disposal, incineration, 136 Pesticide half-lives, soil concentration effect, 127 Pesticide incineration efficiency, 136 Pesticide incineration, incinerator types, 139 Pesticide incineration, muffle furnace temperatures, 137 Pesticide incineration, products of incomplete combustion, 138 Pesticide incineration, temperatures of complete combustion, 137 Pesticide mixtures, disposal problems, 126 Pesticide persistence at high concentrations, 127 Pesticide plastic and paper packaging, incineration, 143 Pesticide unused sprays, treatment & disposal, 144 Pesticide waste constituents, characterization, 132 Pesticide waste, disposal & degradation, 123 ff. Pesticide waste disposal, Australian regulations, 129 Pesticide waste disposal, dark or photoassisted Fenton’s reagent, 158 Pesticide waste disposal, dehalogenation, 151 Pesticide waste disposal, hydrolysis, 150 Pesticide waste disposal, IUPAC recommendations, 179 Pesticide waste disposal, oxidation, 152 Pesticide waste disposal, ozonation, 152, 154 Pesticide waste disposal, photocatalytic oxidation with titanium dioxide, 160 Pesticide waste disposal, photolysis, 150 Pesticide waste disposal, recycling, 129 Pesticide waste disposal, regulatory constraints, 128
206
Index
Pesticide waste disposal, selecting options, 133 Pesticide waste disposal, strategy, 131 Pesticide waste disposal, U.S. regulations, 128 Pesticide waste management methods, Australia, 132 Pesticide waste, unique characteristics, 125 Pesticide wastewater, carbon & lignocellulosic sorption, 146 Pesticide wastewater disposal, 143 Pesticide wastewater, evaporative bed disposal, 145 Pesticide wastewater, options for disposal, 145 Pesticide wastewater, ozonation/bioreactor for cleaning, 148 Pesticide wastewater, sorption/composter for cleaning, 148 Pesticides, chlorinated, destruction by potassium polyethylene glycol ether (KPEG), 140, 142 Pesticides, disposal of unused stocks, 134 Pesticides, not recommended for incineration, 139 pH effects, phosphate compounds rhizosphere, 29 Phosphate amendment, lead-contaminated soil remediation, 4 Phosphate, ammonium, fertilizers, 10 Phosphate compounds, acidity equivalent in soils, 28 Phosphate compounds, antagonism effect on mycorrhizal plant infection, 34 Phosphate compounds, effects on heavy metals in soils, 12, 13 Phosphate compounds, heavy metal immobilization in soils, 18 Phosphate compounds, heavy metal mobilization in soils, 18 Phosphate compounds, heavy metals source in soils, 12 Phosphate compounds, liming value in soils, 28 Phosphate effects on heavy metals, rhizosphere, 29 Phosphate fertilizers, equilibrium dissolution reactions, 11
Phosphate fertilizers, fast-release, 10 Phosphate fertilizers, heavy metal concentrations & sources, 18 Phosphate fertilizers, heavy metal major source, 5 Phosphate fertilizers, slow-release, 10 Phosphate, monoammonium fertilizers, 10 Phosphate, monocalcium fertilizers, 10 Phosphate rocks, soil fertilizers, 10 Phosphate-induced heavy metal adsorption, 22 Phosphates, reactions in soils, 10 Phosphates, water-insoluble compounds, 10 Phosphates, water-insoluble, effects mycorrihizal plants, 32 Phosphates, water-soluble forms, 10 Phosphorus fertilizers, sources of heavy metals in soils, 4 Phosphorus, role in heavy metal availability to soil-plant system, 1 ff. Phosphorus-metal interactions in plants, 20 Photoassisted Fenton’s reagent, pesticide waste disposal, 158 Photohydrolysis, methyl bromide, 60 Photolysis, pesticide waste disposal, 150 Physical/chemical properties, methyl bromide, 49 Phytoremediation, pesticide candidates, 171 Phytoremediation, pesticide contaminated soils, 169 Phytotoxicity, methyl bromide, 66 Plant nutrition, essential elements defined, 2 Potassium polyethylene glycol ether (KPEG), destruction of chlorinated pesticides, 140, 142 Products of incomplete combustion, pesticides, 138 Recycling, pesticide containers, 142 Rhizosphere, acidification by phosphate compounds, 29 Rhizosphere, mycorrhizal effects metal tolerance in plants, 31 Rhizosphere, phosphate-heavy metal effects, 29
Index Sechura phosphate rock, fertilizer, 20 Simulated environmental fate, methyl bromide, 70 Sludge treatment of soils, shifts metals to more mobile forms, 9 Soil adsorption capacity, dictated by cation-exchange capacity, 6 Soil fumigant, methyl bromide environmental fate, 45 ff. Soils, pesticide contaminated cleanup, 161 Soils, source of heavy metals in food crops, 2 Strengite, phosphate fertilizer, 11 Sulfide, methyl bromide degradation in soil, 59 Superphosphate fertilizers, soil reactions, 10 Superphosphate single-, fertilizers, 10 Superphosphate triple-, fertilizers, 10 Temperatures of complete pesticide combustion, 137 Theoretical profile shape method, methyl bromide volatilization, 86 Titanium dioxide photocatalytic oxidation, pesticide waste disposal, 160
207
Trajectory simulation, methyl bromide volatilization, 87 Transport model, methyl bromide, 70, 74 Tricalcium phosphate, fertilizer, 11 Universal Waste Rule, pesticide disposal, 129 Variscite, phosphate fertilizer, 11 Vivianite, phosphate fertilizer, 11 Volatilization boundary condition, methyl bromide, 71 Waste disposal, pesticide, 123 ff. Waste treatment of soils, shifts metals to more mobile forms, 9 Wastewater disposal, pesticide contaminated, 143 Water solubility, metal phosphate compounds, 25 Water-insoluble phosphates, effects mycorrihizal plants, 32 Zero-valent iron treatment, pesticide contaminated soils, 164 Zinc, soil contaminant sources, 3
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Abbreviations min minute(s) M molar mon month(s) ng nanogram(s) nm nanometer(s)(millimicron) N normal no. number(s) od outside diameter oz ounce(s) ppb parts per billion (µg/kg) ppm parts per million (mg/kg) ppt parts per trillion (ng/kg) pg picogram(s) lb pound(s) psi pounds per square inch rpm revolutions per minute sec second(s) sp gr specific gravity sq square (as in “square m”) vs versus wk week(s) wt weight yr year(s)
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