Advances in Agronomy, Volume 104

V O LU M E O N E ADVANCES H U N D R E D IN AGRONOMY F O U R ADVANCES IN AGRONOMY Advisory Board PAUL M. BERTSCH...

Author: Donald L. Sparks

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V O LU M E

O N E

ADVANCES

H U N D R E D

IN

AGRONOMY

F O U R

ADVANCES IN AGRONOMY Advisory Board

PAUL M. BERTSCH

RONALD L. PHILLIPS

University of Kentucky

University of Minnesota

KATE M. SCOW

LARRY P. WILDING

University of California, Davis

Texas A&M University

Emeritus Advisory Board Members

JOHN S. BOYER

KENNETH J. FREY

University of Delaware

Iowa State University

EUGENE J. KAMPRATH

MARTIN ALEXANDER

North Carolina State University

Cornell University

Prepared in cooperation with the American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America Book and Multimedia Publishing Committee DAVID D. BALTENSPERGER, CHAIR LISA K. AL-AMOODI

CRAIG A. ROBERTS

WARREN A. DICK

MARY C. SAVIN

HARI B. KRISHNAN

APRIL L. ULERY

SALLY D. LOGSDON

V O LU M E

O N E

ADVANCES

H U N D R E D

F O U R

IN

AGRONOMY EDITED BY

DONALD L. SPARKS Department of Plant and Soil Sciences University of Delaware Newark, Delaware, USA

AMSTERDAM • BOSTON • HEIDELBERG • LONDON NEW YORK • OXFORD • PARIS • SAN DIEGO SAN FRANCISCO • SINGAPORE • SYDNEY • TOKYO Academic Press is an imprint of Elsevier

Academic Press is an imprint of Elsevier 525 B Street, Suite 1900, San Diego, CA 92101-4495, USA 30 Corporate Drive, Suite 400, Burlington, MA 01803, USA 32 Jamestown Road, London, NW1 7BY, UK Radarweg 29, PO Box 211, 1000 AE Amsterdam, The Netherlands First edition 2009 Copyright # 2009 Elsevier Inc. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means electronic, mechanical, photocopying, recording or otherwise without the prior written permission of the publisher Permissions may be sought directly from Elsevier’s Science & Technology Rights Department in Oxford, UK: phone (+44) (0) 1865 843830; fax (+44) (0) 1865 853333; email: [email protected]. Alternatively you can submit your request online by visiting the Elsevier web site at http://elsevier.com/locate/permissions, and selecting Obtaining permission to use Elsevier material Notice No responsibility is assumed by the publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. Because of rapid advances in the medical sciences, in particular, independent verification of diagnoses and drug dosages should be made ISBN: 978-0-12-374820-1 ISSN: 0065-2113 (series) For information on all Academic Press publications visit our website at elsevierdirect.com Printed and bound in USA 09 10 11 12 10 9 8 7 6 5 4 3 2 1

CONTENTS

Contributors Preface

1. Advances in Assessing Bioavailability of Metal(Loid)s in Contaminated Soils

vii ix

1

Kirk G. Scheckel, Rufus L. Chaney, Nicholas T. Basta, and James A. Ryan 1. Introduction 2. Metal Risks in Soil 3. Biological Metal Uptake 4. Metal Extractability to Predict Availability 5. Metal Chemistry 6. Understanding Metal Bioavailability, Bioaccessibility, and Speciation 7. Conclusions Acknowledgments References

2. Nitrogen in Rainfed and Irrigated Cropping Systems in the Mediterranean Region

3 7 12 16 22 28 41 43 43

53

John Ryan, Hayriye Ibrikci, Rolf Sommer, and Ann McNeill 1. Introduction 2. Mediterranean Agroecosystems 3. Perspective on Nitrogen in Agriculture 4. Fertilizer use Trends in the Mediterranean Region 5. Response of Rainfed Crops to Nitrogen Fertilizer 6. Assessing Soil Nitrogen Status for Crop Yields 7. Nitrogen Fixation Under Mediterranean Dryland Conditions 8. Potential Losses of Nitrogen in Dryland Cropping 9. Integrated Cropping Systems: Implications for Nitrogen 10. Nitrogen in Supplemental Irrigation Systems 11. Nitrogen Tracer use in Rainfed Cropping Systems 12. Modeling of Nitrogen in Rainfed Cropping Systems 13. Future Perspective Acknowledgments References

55 57 63 66 68 80 85 92 95 105 107 113 120 121 121 v

vi

Contents

3. Biogeochemical Processes Controlling the Fate and Transport of Arsenic: Implications for South and Southeast Asia

137

Scott Fendorf and Benjamin D. Kocar 1. Introduction 2. Arsenic Aqueous Chemistry 3. Arsenic Surface and Solid Phases 4. Desorption of Arsenic in Soils and Sediments 5. Biogeochemical Processes 6. Processes Controlling Arsenic Concentrations in South(east) Asia 7. Summary and Conclusions Acknowledgments References

4. Inorganic and Organic Constituents and Contaminants of Biosolids: Implications for Land Application

138 139 139 146 149 154 157 158 158

165

R. J. Haynes, G. Murtaza, and R. Naidu 1. Introduction 2. Sewage Treatment Processes 3. Composition of Biosolids 4. Nutrient Content and Release 5. Heavy Metal Contaminants 6. Organic Contaminants 7. Synthesis and Conclusions References Index

166 168 169 175 182 208 234 237 269

CONTRIBUTORS

Numbers in Parentheses indicate the pages on which the authors’ contributions begin.

Nicholas T. Basta (1) School of Environment and Natural Resources, The Ohio State University, Columbus, Ohio, USA Rufus L. Chaney (1) USDA-ARS, Environmental Management and Byproduct Utilization Laboratory, Beltsville, Maryland, USA Scott Fendorf (137) Stanford University, Stanford, California, USA R. J. Haynes (165) School of Land, Crop and Food Sciences/CRC CARE, The University of Queensland, St Lucia, Australia Hayriye Ibrikci (53) Soil Science Department, Faculty of Agriculture, C ¸ ukurova University, Balcali, Adana, Turkey Benjamin D. Kocar (137) Stanford University, Stanford, California, USA Ann McNeill (53) Adelaide University, Roseworthy Campus, Adelaide, South Australia, Australia G. Murtaza (165) Centre for Environmental Risk Assessment and Remediation, Division of Information Technology, Engineering and the Environment, University of South Australia, Mawson Lakes Campus, South Australia, Australia and Institute of Soil and Environmental Sciences, University of Agriculture, Faisalabad, Pakistan R. Naidu (165) CRC CARE, Salisbury, South Australia, Australia James A. Ryan (1) USEPA, Cincinnati, Ohio, USA John Ryan (53) International Center for Agricultural Research in the Dry Areas (ICARDA), Aleppo, Syria vii

viii

Contributors

Kirk G. Scheckel (1) USEPA, National Risk Management Research Laboratory, Cincinnati, Ohio, USA Rolf Sommer (53) International Center for Agricultural Research in the Dry Areas (ICARDA), Aleppo, Syria

PREFACE

Volume 104 contains four outstanding reviews on timely topics that will be of interest to plant, soil, and environmental scientists. Chapter 1 is a comprehensive review on frontiers in assessing the bioavailability of metal (loids) in contaminated soils. Topics that are covered include metal risks in soils, biological metal uptake, metal chemistry, metal extractability and prediction of availability, and advances in understanding metal bioavailability, bioaccessibility, and speciation. Chapter 2 discusses nitrogen in rainfed and irrigated cropping systems in the region including Mediterranean agroecosystems, fertilizer use trends, response of rainfed crops to nitrogen fertilizer, nitrogen fixation under Mediterranean dryland conditions, and modeling of nitrogen in rainfed cropping systems. Chapter 3 covers the biogeochemical processes that impact the fate and transport of arsenic with specific emphasis on South and Southeast Asia. Processes that are critical include ion displacement, desorption, reduction of arsenate to arsenite, and reductive dissolution of Fe- and Mn-(hydr)oxides. Chapter 4 provides a thorough treatment on inorganic and organic contaminants in biosolids and impacts on application to land. Discussions on sewage sludge treatment processes, nutrient content and release, heavy metal contaminants, and organic contaminants are provided. The authors are congratulated on their first-rate reviews. DONALD L. SPARKS Newark, Delaware, USA

ix

C H A P T E R

O N E

Advances in Assessing Bioavailability of Metal(Loid)s in Contaminated Soils Kirk G. Scheckel,* Rufus L. Chaney,† Nicholas T. Basta,‡ and James A. Ryan§,1 Contents 3 7 7 8 11 12 12 13 14 16 16

1. Introduction 2. Metal Risks in Soil 2.1. Bioavailability and soil element risks 2.2. Phytotoxicity risks from soil elements 2.3. Risks to soil organisms 3. Biological Metal Uptake 3.1. Risks through soil ingestion 3.2. How much soil do children ingest? 3.3. Food-chain transfer and risks 4. Metal Extractability to Predict Availability 4.1. In vitro bioaccessibility 4.2. Common soil extractions to predict risk of phytotoxicity or food-chain risk 5. Metal Chemistry 5.1. Metal equilibrium in soils 5.2. Metal speciation in soils 6. Understanding Metal Bioavailability, Bioaccessibility, and Speciation 6.1. Lead 6.2. Arsenic 7. Conclusions Acknowledgments References

* {

{ } 1

21 22 22 25 28 28 35 41 43 43

USEPA, National Risk Management Research Laboratory, Cincinnati, Ohio, USA USDA-ARS, Environmental Management and Byproduct Utilization Laboratory, Beltsville, Maryland, USA School of Environment and Natural Resources, The Ohio State University, Columbus, Ohio, USA USEPA, Cincinnati, Ohio, USA Retired.

Advances in Agronomy, Volume 104 ISSN 0065-2113, DOI: 10.1016/S0065-2113(09)04001-2

#

2009 Elsevier Inc. All rights reserved.

1

2

Kirk G. Scheckel et al.

Abstract The term bioavailability has many different meanings across various disciplines of toxicology and pharmacology. Often bioavailability is concerned with human health aspects such in the case of lead (Pb) ingestion by children. However, some of the most contaminated sites are found in nonpublic access facilities (Department of Defense or Energy) or in remote regions as a result of mining or industrial practices in which ecoreceptors such as plants, animals, and soil organisms are the primary concerns as well as the potential for food-chain transfer. In all cases, the endpoint requires movement of the element across a biological barrier. The still utilized approach to base risk assessment on total metal content in soils is an outdated endeavor and has never been proved to be scientifically sound. Yet to reverse this trend, much work is required to establish baseline bioavailability measurements and to develop complementary methods that are capable of predicting bioavailability across a whole range of impacted media in a cost-efficient manner. Thus, regulators have recognized site-specific human health risk assessments play a key role in decision-making processes at contaminated sites. Bioavailability issues surrounding metal-contaminated soils and media have been an area of intense research. For obvious ethical reasons, we cannot solicit humans, in particular the sensitive population of children, from the general population for experimental purposes to examine the long-term harmful effect of metals in soils. However, some adult human feeding studies have been accomplished under tight medical supervision and with very small doses. One option to understand and relate bioavailability in humans is to employ animal surrogates; however, the physiology of most animals is different than that of humans but good correlations have been achieved despite the dose–response paradigm not being identical. The biggest drawback of in vivo studies to examine metal bioavailability to an appropriate ecoreceptor, be it human, plant, or soil organism, is the tremendous cost and time involved relative to chemical and physical surrogates. Chemical surrogate methods generally only require knowledge of the total metal content so that a percent bioaccessible number can be generated from in vitro extractions that simulate digestive systems or mimic responses to sensitive ecoreceptors. However, there is not a consensus as to which of the many in vitro methods is the best analogy to an ecoreceptor uptake and the same can be said for in vivo animal models to mimic human response as well. Further, there is yet to be a single in vitro method that can account for more than a few elements for a specific exposure pathway (e.g., Pb and/or arsenic (As) for human health). These in vitro tests require honest and accurate validation against in vivo bioavailability measurements, but most of all would benefit from metal speciation methods to identify the forms of metals allowing their release. Adaptation of spectroscopic speciation techniques to identify metal(loid) phases is extremely beneficial in bioavailability research to understand the variability of biologically available metal uptake, to manipulate the ecosystem to reduce bioavailability via in situ amendments, to monitor the long-term stability of elements to ensure bioavailability indicators do not change over time, and to develop comprehensive predictive models based on speciation.

Advances in Assessing Bioavailability of Metal(Loid)s in Contaminated Soils

3

1. Introduction In most cases, the toxicity of contaminants depends on how much of it is absorbed into the body or taken up by plants. For soil contaminants where human exposure is by ingestion of soil or plants and organisms produced on the soil, toxicity depends on absorption into the gastrointestinal (GI) system. Information on how well a contaminant is absorbed into the GI system is important to determining how much of a contaminant humans can be exposed to before health effects occur. Because typical health effect dose– response assessments (and resulting oral reference doses (RfDs) and cancer slope factors (CSFs)) are generally expressed in terms of ingested dose (rather than absorbed dose into the organism), accounting for potential differences in absorption between different exposure media can be important to site risk assessment (USEPA, 1989). Thus, if the oral RfD for a particular metal is based on bioavailability studies in water, risks from ingestion of the metal in soil or plant produced on the soil might be (likely is) overestimated. Minor adjustments in oral bioavailability based on nonrelevant exposure pathways can have significant impacts on estimated risks and cleanup goals for hazardous waste sites (USEPA, 1989). It is increasingly recognized that the response of an at-risk population is not controlled by the total metal concentration, but instead is controlled by only the biologically available portion, which is dependent on the route of exposure, the pharmacokinetics of the organism, and the speciation of the contaminant. In spite of the earlier understanding, the complexity of metal-contaminated sites has and continues to be simplified to a measure of the total metal content. Regulations on the fate and effects of metals in the environment based solely on total concentrations are no longer (perhaps never has been) valid, state-of-the-art, or scientifically sound. A vast amount of knowledge clearly illustrates the decisive role of metal speciation when metal bioavailability and phytoavailability in the environment have to be assessed (McNear et al., 2007; Ryan et al., 2004). While total metal content is a critical regulatory measure in assessing risk of a contaminated site, total metal content alone does not provide predictive insights on the bioavailability, mobility, and fate of the metal contaminants. Thus, a better understanding of the nature of the chemical and physical interactions of contaminants with soil constituents can increase the scientific understanding and lead to regulatory and public confidence in the use of bioavailability adjustments. Predictions of long-term stability rely on a mechanistic understanding of how contaminants are stored or sequestered within the soil. Bioavailability processes are defined as the individual physical, chemical, and biological interactions that determine the transfer of chemicals associated with soils to plants and animals. Bioavailability processes are

4


embedded within existing human health and ecological risk frameworks to reduce uncertainty in exposure estimates and improve risk assessment (USEPA, 2007b). In both ecological and human health risk assessment, bioavailability is usually reflected in default values or site-specific data that are inserted into exposure equations. Although a multitude of processes can affect bioavailability, a typical risk assessment generates one value that is used to adjust the applied dose. For this reason, many bioavailability processes are hidden within risk assessment, and assumptions made about these processes are sometimes not clear. Although long employed in toxicology and agricultural sciences, the concept of bioavailability has recently sparked the interest of the hazardous waste industry as an important consideration in deciding how much waste to clean up. This interest stems from observations that some contaminants in soils appear to be less available to cause harm to humans and ecological receptors than is suggested by their total concentration, such that cleanup levels expressed as total concentrations poorly correlate with actual risk. Correct characterization of bioavailability in contaminated soils and sediments may indicate that greater levels of contamination can be left untouched without increased risks, thus, reducing cleanup costs and reducing volumes of contaminated media requiring intrusive remedial options (USEPA, 2007c). However, in order to pursue this concept in risk assessment critical knowledge of bioavailability processes and spectroscopic speciation techniques are required to develop a mechanistic understanding of the bioavailability processes to improve the science of risk assessment to develop predictive models derived from sound research. Further, chemical, environmental, and regulatory factors must align in considering bioavailability processes that influence risk-based decision-making (NRC, 2003). Because the fraction of a soil element which can actually be absorbed by an organism to cause harm depends on the chemical forms present and physical/chemical properties of the soil, in both risk assessment and remediation evaluation, the fraction of a soil element which can actually cause harm must be identified. This fraction is ultimately defined as the bioavailable fraction, and because measurement of the bioavailable fraction is timeconsuming and expensive via in vivo animal feeding studies, in vitro chemical methods are being developed to estimate the bioavailable fraction. In the case of ingestion of soil, the in vitro or chemical estimation method has been labeled ‘‘bioaccessible’’ (to avoid confusion with ‘‘bioavailable’’) and is a measure of the amount of metal that can be liberated from the soil matrix, thus not a measure of the amount of metal that moves across the GI epithelium to harm internal target tissues and organs. Extensive progress has been made in development of soil Pb and As bioavailability testing in conjunction with bioaccessibility methods (Drexler and Brattin, 2007; Rodriguez et al., 2003; Ruby et al., 1993, 1996). Additionally, great effort has been wasted in planning inconsequential research efforts to develop


5

bioaccessibility methods, which try to match all digestion processes without a valid bioavailability endpoint as a comparison. In the end, an in vitro bioaccessibility method only needs to be well correlated with an acceptable in vivo bioavailability model. Actually, the simpler and less expensive the bioaccessibility method can be made, the better, as long as the correlation with bioavailability is high. Further, it is necessary for the tests to be reproducible in laboratories across the globe, which has not been the case for many of the bioaccessibility methods available today. Further, for such methods to be relevant to testing of remediation methods, changes in bioavailability due to field treatments should be reflected in the bioaccessibility test results. In the case of soil Pb, in situ remediation using phosphate and other treatments have been proved to reduce bioavailability to pigs, rats, and humans, but the bioaccessibility test conducted at pH 1.5 does not measure this 69% reduction in bioavailability to human adults while testing at pH 2.2 or 2.5 does reflect the effectiveness of the soil treatment (Ryan et al., 2004). Other simple chemical tests have been shown to suffer significant flaws in that the extraction causes changes in chemical speciation during the test, and have not been shown to correlate with bioavailability changes due to soil treatments. Further, it is necessary to have a valid measure of why the bioavailability or bioaccessibility of samples are different and whether the changes are persistent; thus, the need for metal speciation. For sensitive ecoreceptors (plants, animals, and soil organisms), where testing with the organism to be protected is more readily conducted, chemical methods have been developed which integrate potential toxicity across soil properties including pH which often strongly affects bioavailability. Mild neutral salt extractions (similar to the first extraction step of a sequential extraction procedure) are often found to be effective methods. However, assessment of potential toxicity by adding metal salts to uncontaminated soils substantially fails to mimic field contaminated soils because elements react with soils, and metal salt additions alter soil pH and do not account for the aging effect of metals in soils. Traditional toxicology approaches of adding element salts and immediately measuring toxicity are clearly inappropriate, and can cause serious artifacts due to pH change resulting from the metal salt addition, or formation of soluble metal complexes which temporarily increase or decrease element bioavailability. Thus, testing of potential toxicity has as many problems as testing of bioaccessibility. It seems clear that by taking present knowledge into account, effective toxicity testing, bioaccessibility evaluation, and risk assessment can provide massive savings to the public in dealing with contaminated soils. The extent to which metals are bioavailable has significant implications on human and ecological health following exposure and on potential remediation of contaminated sites. Characterization via speciation of insoluble metal phases in contaminated soils and sediments may indicate that greater levels of contamination can be left untouched without increased risks, thus,

6


driving reduced cleanup costs and limited volumes of contaminated media through less intrusive remedial options. A mechanistic understanding of the bioavailability process in relation to metal speciation will allow development of predictive models and improvement of risk assessment. Further, chemical, environmental, and regulatory factors must align in considering bioavailability processes that influence risk-based decision-making (NRC, 2003). In both ecological and human health risk assessment, bioavailability is usually reflected in default values or site-specific data that are inserted into exposure equations. Although a multitude of processes can affect bioavailability, a typical bioavailability assessment generates one value that is used to adjust the applied dose. The Risk Assessment Guidance for Superfund, Volume I: Human Health Evaluation (Part A) (RAGS) (USEPA, 1989) supports the consideration of bioavailability in the determination of site-specific human health and environmental risks. This guidance has been used to support bioavailability adjustments across different routes of exposure at contaminated sites. However, the use of bioavailability information in site-specific risk assessment has not been widespread (due to limited data, uncertain methodologies, and lack of method validation). The primary impediment to the broad use of bioavailability data in risk assessment and decision-making is the absence of rapid and inexpensive tools that can generate reliable relative bioavailability (RBA) estimates in the receptors of concern. It is in this context that coupling in vivo bioavailability, in vitro bioaccessibility, and speciation research can fill many data gaps to aid in understanding and predicting bioavailability. The speciation, or chemical form, of metals governs their fate, toxicity, mobility, and bioavailability in contaminated soils, sediments, and water. Different chemical forms of metals, for example, can differ greatly in the amounts taken up by organisms. The varying bioavailability values of different metal species is a large reason for the wide range of bioaccessibility values measured using standardized in vitro analyses of different soils. Other interactions between metals and soil components also govern speciation and affect bioavailability. The influence of the soil matrix on metal(loid) availability is in constant dynamic equilibria with multiple independent variables such as solid mineral phases, exchangeable ions and surface adsorption, nutrient uptake by plants, soil air, organic matter, and microorganisms, and water flux. However, determining speciation is not a trivial task, particularly at low concentrations in a complex matrix such as soil. To assess these chemical properties and to accurately gauge their impact on human health and the environment we need to characterize metals at the atomic level with spectroscopic techniques. This research must move beyond operationally defined sequential extraction methods and utilize analytical instruments that are capable of identifying metal species (D’Amore et al., 2005) Researchers have used advanced synchrotron radiation methods to elucidate the true, in situ speciation of metal contaminants. Synchrotron techniques include X-ray absorption near-edge spectroscopy (XANES), which identifies the


7

oxidation state and first coordination shell and X-ray absorption fine structure (XAFS) spectroscopy provides information on coordination environment of a selected element as well as interatomic bond distances and identity of nearest-neighboring atoms to determine speciation. These methods can also be used in conjunction with statistical methods (principal component analysis and linear combination fitting) to determine chemical phases via a finger printing process with a library of known reference standards. Although most soil criteria and regulations for metals are still based on the total concentration of the metal in question, it is becoming more and more evident that spectroscopic speciation is vital for regulatory risk assessment of environmentally relevant metals in conjunction with in vivo animal data and validated in vitro extractions for human health effects and plant uptake/foodchain transfer for sensitive ecoreceptors. These innovative research tools are expanding our ability to directly identify the role of metal speciation on many dynamic processes that influence bioavailability and risk. The application of synchrotron techniques for the speciation of metals to assess bioavailability seems logical to this chapter, but to the common regulator a simpler approach has been to pick a fractional number relative to the total concentration of a metal in order to establish a cleanup standard. Fortunately from a human health perspective, the common regulator approach is significantly conservative almost to a point that hinders common sense for site remediation. A good example of this is arsenic which is regulated assuming 100% bioavailability, yet several studies have demonstrated that absolute bioavailability of arsenic at most sites can be as low as 20% through matrix effects or natural attenuation processes. If a lower bioavailability value can be utilized at a site, then the effective cleanup standard is raised resulting in significant savings in remedial clean up costs without harm to human health or the natural environment. However, few speciation studies have truly taken on the task of addressing bioavailability from start to finish—meaning many synchrotron-based studies will broadly state that their results support an understanding or prediction of bioavailability but provide no real data on bioavailability to support the claim. There is much speciation research needed to complement in vivo and in vitro research on metal bioavailability that can lead to effective predictive models on the long-term fate of contaminants.

2. Metal Risks in Soil 2.1. Bioavailability and soil element risks The focus of this chapter is on the potential for adverse effects of soil elements to organisms; specifically soil organisms, plants, livestock, wildlife, and humans which ingest soils and crops grown on soil. The most common

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understanding of bioavailability of a soil element is the fraction of total soil element which can be absorbed into an organism and cause an adverse or beneficial effect in the exposed organism. In its concern with direct ingestion of soil, the USEPA has defined bioavailability as the fraction of an ingested dose that crosses the GI epithelium and becomes available for distribution to internal target tissues and organs (USEPA, 2007b). From this definition, bioavailability can be divided into two kinetic steps: (1) dissolution and liberation of the metal in GI fluids and (2) absorption of the metal across the GI epithelium into the blood stream. Either of which can be ratelimiting to element bioavailability. Combining the variability of geochemical forms of elements in contaminated soils with dissolution chemistry and biological absorption processes in the GI tract is a complex endeavor but should be a call to arms for the many researchers pursuing this effort. The scientific and regulatory communities must push further convoluting of this complexity by recognizing that each element has its own specific environmental toxicology; meaning the organism to which a specific element can cause an adverse effect at the lowest environmental exposure and the interaction of other factors with that element such as Ca with Pb, Zn with Cd, Fe with As, and Cu with Mo. In some cases, the key interaction which affects element risk is related to dissolution from ingested soil, while in other cases, interaction during intestinal absorption is the key process which controls risk from an element. This understanding must come from assessment of the specific pathway from soil to organism for each element which can harm a sensitive exposed organism. Often children are the most exposed and sensitive organisms with respect to contaminated soils in urban areas, but for remote contaminated areas, it is wildlife, plants, or soil organisms that are likely to be the most exposed and sensitive organisms. But each element has its specific chemistry in soils, potential for uptake by plants or soil organisms, and potential to affect consumers of plants or soils.

2.2. Phytotoxicity risks from soil elements The most sensitive adverse effect of some elements in soils is phytotoxicity. It seems clear that the first limiting effect of Zn, Ni, Cu, Mn, Al, and possibly some other elements are phytotoxicity to sensitive plants. Of course, plant species vary in tolerance of soil elements. And soil properties can strongly affect phytotoxicity. For cations, acidic soil pH strongly promotes element toxicity, and the elements react over time increasingly strongly to lower phytoavailable forms. In the case of Ni, it was shown by Singh and Jeng (1993) that Ni was about 10-fold less accumulated by perennial ryegrass over a 3-year test period using experimental methods which are highly defensible. Initially, such results were explained in terms of adsorption and diffusion into micropores of the sesquioxides (Bruemmer et al., 1988). Since then, research has shown that new mineral phases may


9

form in Ni enriched soils, both Ni–Al layered double hydroxides (LDH) and Ni-silicates (Scheckel and Sparks, 2001). And although Zn can also form such LDH species, the Zn forms are weaker than the Ni forms (Roberts et al., 2003). Cu and Cd apparently do not form the LDH species in soils; however, Co may form them (Scheinost and Sparks, 2000). Figure 1 summarizes known soil reactions of Ni in relation to plant uptake and Ni phytotoxicity. Some industrial compounds can land on soils and persist for long periods. For example, NiO dissolves very slowly, with a half-life of 20.4 years at pH 7.25, and slower with larger particle size (Ludwig and Casey, 1996). A study of Ni species in a smelter-contaminated soil at Port Colborne, Ontario found particles of NiO remaining in the soil more than 30 years after smelting ceased (McNear et al., 2007). They also found that Ni-LDH had formed in these soils over time, confirming the practical significance of Ni-LDH formation in contaminated soils. Ni-sulfides deposited on soils can be oxidized by microbes. Other Ni phases enter into equilibria with soil sorption surfaces and ligands. Much soluble soil Ni2þ is chelated or complexed, but the free ion shuttles among sites based on free energy and binding site specificity. As shown in Fig. 2, grasses suffer an unusual symptom of Ni-induced Fe deficiency chlorosis in which the severity follows a diurnal pattern (banded chlorosis). Phytosiderophores (PSid) are secreted by young grass roots to dissolve soil Fe, and the Fe-PSid is absorbed by a transport protein specific to the Fe-PSid. At low pH Ni fills the PSid and can push Fe out by competition, but during the morning pulse of PSid secretion, some Fe is dissolved and absorbed so part of the growing

Grasses PSid

Chelated to organic matter humics and fulvics

Adsorbed on Fe/Mn oxides

Occluded in Fe/Mn oxides

FeOx

Plant shoots

Plant roots

Soil solution Ni2+ + L

NiL

Inorganic solids Ni(OH)2 NiO

L = ligands organic inorganic Soil microbes

Ni-silicate Ni-Al-LDH

Figure 1 Equilibria of Ni in soils in relation to uptake by both dicots and grasses; note formation of Ni-Al-LDH and Ni-silicate over time which reduces Ni phytoavailability. PSid are phytosiderophores such as deoxymugineic acid secreted by wheat to chelate soil Fe.

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Oat

Barley

Ni toxicity symptoms in oat and barley seedlings.

Figure 2 Unique symptoms of Ni-induced Fe deficiency (Ni-phytotoxicity) with diurnal variation in severity which results from Ni preventing Fe-phytosiderophore formation in the rhizosphere except during morning pulse secretion of phytosiderophore by young grass roots.

leaf blade receives Fe before it emerges from the culm. As pH is raised, Ni is bound increasingly strongly by soil sorbents, and forms new solids (Ni-LDH, Ni-silicates) such that insufficient Ni remains reactive to compete for filling the PSid in the rhizosphere (Kukier and Chaney, 2004). Simply making soils calcareous can remediate Ni phytotoxicity potential for species which are very sensitive at acidic pH (Siebielec et al., 2007). Interestingly, Cu is more bound by organic matter than Fe and Mn oxides, nor does it form LDH compounds in soils, so as pH is raised and Fe is less available for chelation by PSid, Cu inhibits Fe uptake in a simple Fe deficiency pattern (Michaud et al., 2007) rather than the banded chlorosis caused by Ni and Co. Zn forms LDH compounds, and is readily converted to lower phytoavailability forms in soil, so that making a high Zn soil calcareous with reasonable soil fertility remediates Zn phytotoxicity to sensitive plants (Li et al., 2000). Unfortunately, others followed the toxicological approach to establish limits for soil Ni, by adding soluble Ni salts followed by immediate cropping, failing to correct for the metal salt-induced drop in soil pH (Speir et al., 1999) resulting in exaggerated soil solution Ni concentrations (Oorts et al., 2006; Rooney et al., 2007; Thakali et al., 2006). Yet others studied nutrient solutions but did not understand basic metal chelate equilibria in nutrient solutions and observed apparently higher toxicity at higher pH (Weng et al., 2003), in strong contrast with the real world (Kukier and Chaney, 2004; Siebielec et al., 2007). McNear et al. (2007) examined the


11

speciation of Ni in Welland loam and Quarry muck soils around a refinery and relate these findings to Ni mobility and bioavailability. XAFS and X-ray fluorescence (XRF) showed that Ni–Al LDH phases were present in both the limed and unlimed mineral soils, with a tendency toward more stable Ni species in the limed soil, possibly aided by the solubilization of Si with increasing pH. Precipitation of some mineral phases in wetland sediments can potentially limit metal bioavailability through sequestration in low-solubility compounds, such as metal sulfides. Analysis of XAFS data confirmed that sulfide compounds dominated zinc speciation throughout the sediment in a study by Peltier et al. (2003). Uptake of trace metals in Phragmites plants was limited primarily to plant roots, while concentrations of both Pb and Zn in other aquatic vegetation were significantly elevated, representing a potential bioaccumulation hazard and possible food-chain transfer concern for local wildlife. Another example of synchrotron research to understand plant uptake of contaminants was conducted by Punshon et al. (2005). Synchrotron XRF (S-XRF) demonstrated changes in Ni and U distribution in wheat grown on contaminated soil and the distribution of Ca, Mn, Fe, Ni, and U in roots of willow growing on a former radiological settling pond, with U located outside of the epidermis and Ni inside the cortex with confirmation by microtomography. Further, XRF and XANES linked the elevated Se concentrations in sediments of a coal fly ash settling pond at the site with oral deformities of bullfrog tadpoles.

2.3. Risks to soil organisms Toxicity to soil microbes and fauna has received much study, but often the methods used suffered from serious artifacts much as noted earlier for phytotoxicity. Addition of metal salts to soils is even more inappropriate in study of soil organisms because the organism receives the shock of soluble added elements rather than the metals equilibrated with the soil. Complexes of the metals with anions can cause persistence of soluble ions, and high rates of metal cations can drive pH several units lower greatly increasing soil metal solubility. The effects of diverse soil properties on metal toxicity to earthworms are considered by Lanno and Basta (2003), Bradham et al. (2006), and Dayton et al. (2006). In addition, remediation of phytotoxicity is often successful for remediation of toxicity to soil microbes and fauna (Brown et al., 2004, 2005, 2007; Conder et al., 2001). As we have noted, when metals are present at phytotoxic levels, the recommended remediation treatment would be to make the soil calcareous to minimize metal phytoavailability and provide a persistent remediation. Because these treatments give lower and lower metal bioavailability over time, it generally provides effective protection of soil

12


organisms. And consumers of soil organisms appear to be protected except for soil ingestion risks (Pb, As, F) where earthworms can carry a high fraction of dietary soil into diets of earthworm consumers. Risks from soil Cd to earthworms and earthworm consumers have often been overestimated (Brown et al., 2002a,b). In estimating bioaccumulation ratios, one needs to take into account that the ratios are 10-fold higher for background uncontaminated soils than for contaminated soils. Predictions of risks to earthworm consumers have not been confirmed except for the case of a Cu–Cd smelter at Prescott, UK. Because Zn was not present with the Cd, earthworms accumulated high body burdens of Cd without injury that would have occurred from Zn in most contamination cases. In mine waste studies, Cd bioaccumulation was clearly limited by the presence of Zn (Andrews et al., 1984). Tolerance of soil microbes to metals is very complex, and traditional methods of study by adding metal salts to soils clearly confound the tests. Soils with deficient Zn have microbes which are less resistant to Zn additions than found in soils with Zn contamination. These findings led McLaughlin and Smolders (2001) to introduce the concept of ‘‘metalloregion’’ to suggest that some soils may be much more resistant to additions of Zn than other soils; that is, it would be an error to apply results from the most sensitive soil to all soils. Although it is clear that white clover rhizobium is relatively sensitive to excessive soil Zn, it is also very sensitive to simple soil acidity; causation in selection of ineffective nodulating strains was more affected by low soil pH than by soil Cd or Zn levels (Ibekwe et al., 1997). In our experience, sensitive plants are less resistant to excessive bioavailable soil metals than are the microbes in the soil, such that protection against phytotoxicity protects soil function.

3. Biological Metal Uptake 3.1. Risks through soil ingestion For selected elements, the element in ingested soil can comprise a risk to animals or humans and is especially well studied for Pb and As, but also considered important for F, Hg, and other elements. Soil ingestion circumvents the soil-plant barrier whereby limited plant uptake limits significant exposure. In soil ingestion, an element must have sufficient bioavailability/ solubility that it can be absorbed in the intestine to a greater extent than if garden foods growing on the soil were consumed. It has been recognized for decades that Pb deposited on the outside of forages can cause adverse effects in grazing livestock. Then, as risks from Pb in the urban environment were studied in more detail, it became apparent that Pb-rich exterior soil and dust can be carried into homes and provides


13

exposure to young children who do not play outdoors. And that Pb paint dust ingested by hand-to-mouth transfer could be the important pathway for Pb exposure. Additional research eventually showed that interior paint Pb comprised far greater risk than soil Pb (Lanphear et al., 1998). But a key learning was that soil Pb was a greater risk through soil ingestion than through uptake by garden food crops (Chaney and Ryan, 1994). Pb uptake by plants can occur, but uptake of equilibrated soil Pb is small; soil adhering on low-growing crops is a more important source of Pb risk than is Pb uptake by plants. Gardening in urban soils is a difficult issue; if gardeners avoid growing low-growing leafy and root vegetables, and take care to exclude soil from their homes, gardening can be a safe practice until soils exceed levels, which comprise a clear risk by soil/dust ingestion. Soil Pb became a worrisome source of risk to children because Pb has become widely dispersed in urban soils (Mielke et al., 1983, 2007) as well as at industrial and DOD sites. Paint, building demolition dispersing interior paint (Farfel et al., 2005a,b), stack emissions, and automotive exhaust emissions contributed to urban soil Pb loadings. Center city soils are considerably more contaminated than suburban soils, although exterior Pb-paint scrapped to soil can cause massive soil contamination wherever it occurs, easily causing soil to exceed 10,000 mg Pb kg 1.

3.2. How much soil do children ingest? Several studies have been conducted to estimate soil ingestion by young children. Some investigators measured soil on hands of children and how long after starting play it took for their hands to become contaminated. The most widely accepted estimate of chronic soil ingestion by young children were reported by the team of Calabrese, Barnes, and Stanek at University of Massachusetts. They used ICP-AES and later ICP-MS to measure tracer elements in feces of children recovered from diapers. They analyzed diets to allow correction for dietary intake of elements, and provided toothpaste low in Ti so that fecal Ti might measure soil. Over time, they discovered that some of the elements they originally used as tracers were present at lower levels in the fine soil fraction ( 1) then there was a potential enhancement of metal bioavailability due to insufficient coprecipitation of the metals with P FeS. However, the results demonstrated that SEM/AVS was not a valid assumption P except in the case of Cd. Nontoxic responses were noted when SEM/AVS ranged from 2.7 to 5.25 and 100% toxicity was P documented when SEM/AVS was equal to 0.54 indicating that FeS was not the primary contributor to AVS extraction results. XAS speciation found that metal contaminants were present in the sediment matrix as both sulfide and oxide solid phases invalidating the assumption that the metals were controlled by iron monosulfide partitioning in sediments—a very P important point considering the many assumptions behind the SEM/AVS approach. This study illustrated the need for multiple tests to assess bioavailability processes in ecological receptors and that spectroscopic techniques can verify contaminant speciation to provide a molecular understanding for interpreting toxicity test results. Probably, the most comprehensive field study to link synchrotron metal speciation to human bioavailability was the research of Ryan et al. (2004) in the examination of Pb-contaminated soil treated with phosphate amendments to reduce Pb bioavailability as part of the USEPA’s In-Place Inactivation Natural Ecological Restoration Technologies (IINERT) Soil-Metals Action Team. In the study, Pb speciation via XAFS and XRF mapping, in vitro extraction at three solution pH values, and in vivo animal feedings to swine, rats, and humans were conducted on 12 different treatments at three

28


different aging times. The objective of the study was to convert soil Pb to pyromorphite, a Pb phosphate mineral, which is extremely insoluble and, thus, presumed nonbioavailable. A linear trend was observed when comparing the amount of pyromorphite present in the soil samples determined by XAFS to the amount of bioaccessible Pb measured by the in vitro extraction method. It was demonstrated that in situ treatment can reduce soil lead bioavailability and this conclusion was also demonstrated for the in vitro data and its relationship to synchrotron speciation. Therefore, the study illustrated that the conversion of soil Pb to pyromorphite can be quantified by synchrotron speciation techniques to indicate increasing pyromorphite concentration translates into decreasing in vitro and in vivo Pb bioavailability. Further details on this study are discussed in the following sections.

6. Understanding Metal Bioavailability, Bioaccessibility, and Speciation 6.1. Lead 6.1.1. How much soil Pb is too much? Greatest risk from soil Pb depends on getting the soil into the area where it can be ingested by hand-to-mouth play and exploration by children. Growing children are very sensitive to excessive absorbed Pb, and absorb a higher fraction of dietary Pb than do adults. Epidemiologic studies in Pb-dust contaminated housing show that peak blood Pb in childhood occurs at about 18–24 months age, but that is still before children are allowed to play unsupervised in soil. Therefore, interior dust, paint dust, and soil/dust brought into the house must provide the Pb exposure which causes the bulk of excessive soil Pb absorption by children. This process was first proved when the clothing of Pb workers raised Pb levels in house dust and caused Pb poisoning of their children even though their housing did not have high Pb paints (Dolcourt et al., 1978). Although the focus here is on soil Pb, it must be recognized that paint Pb in a home is the much more likely source of high Pb levels in house dust and excessive Pb absorption in children. In several smelter town studies, for at least some of the children the majority of blood Pb came from paint rather than from soil or industrial dusts (Gulson et al., 2004). In the same token, industrial dusts emitted from smelters, or resuspended Pb-contaminated dust in an arid community readily recontaminates household dust and remain key sources of excessive blood Pb. In Trail, British Columbia, it was found that blood Pb dropped substantially when a new flash smelting technology was introduced which caused much lower Pb emissions, with a corresponding drop in house dust Pb (Hilts, 2003).


29

For Pb absorption to occur, the chemical forms of Pb in the soil/dust must be absorbable when the soil is ingested. Research has shown that some Pb minerals are poorly absorbed by humans (PbS, chloropyromorphite (CP)), while some others are readily absorbed (PbCO3). Perhaps the most important factor in Pb absorption is the presence of food in the stomach/ intestine when the Pb source is present. Several research teams evaluated Pb isotope absorption by human volunteers fed Pb with meals or specific foods, or on fasting. On fasting, soluble Pb is absorbed at 50–80%, usually assumed by EPA to be 50% for Pb acetate. But when the Pb is ingested with a meal, 1 h before a meal, or up to 4 h after a meal, absorption falls to the range of 2–5% of dose (Heard and Chamberlain, 1982; Heard et al., 1983; James et al., 1985). Particular food ingredients ingested with Pb can greatly reduce Pb absorption, especially Ca (Blake and Mann, 1983); Ca is believed to compete with Pb absorption by a Ca-transport protein in the small intestine, but also to form coprecipitates with phosphate and Pb. Pb incorporated in kidney or spinach had quite low bioavailability to adults (Heard et al., 1983). One key learning in this area of science came from a test of soil removal and replacement. The USEPA conducted a congressionally supported test in three cities of whether removal and replacement of soil would cause a reduction in blood Pb in the children who lived there. The biokinetics of blood Pb in children versus exposure was considered, and it was believed that if soil Pb absorption by children was reduced for a year by the soil replacement, blood Pb should decline significantly if soil were contributing to the Pb which was being absorbed. Tests were conducted in three cities. Children who lived in areas where soils around houses contained at least 500 mg Pb kg 1 were identified and volunteers were assigned to early replacement and late replacement (at the end of 1 year, the second half of the population would have their soil replaced). Blood Pb was sampled before any changes were made, after the first half were replaced, and several more times until the end of the test. The children were randomly (of the general area where they lived) assigned to early or late replacement so it was a randomized test. In Baltimore (Farrell et al., 1998) and Cincinnati, soils which were replaced contained only about 500 mg Pb kg 1, and there was no significant reduction in blood Pb in the children. In Boston, children were assigned to early or late replacement as they joined the study; the soils replaced contained between 1800 and 2000 mg Pb kg 1; soil replacement included dust control post soil replacement to assure full removal of the exposure source that may have been stirred up during replacement; replacement plus dust control was compared with dust control alone versus absolute control, and soil replacement gave a small significant reduction in blood Pb at this high soil Pb level (Weitzman et al., 1993). The ‘‘late’’ replacement also gave a small significant reduction in blood Pb (Aschengrau et al., 1994): ‘‘The combined results from both phases suggest that a soil lead reduction of 2060 ppm is associated with a 2.25–2.70 mg dL 1 decline in

30


blood lead levels’’ (from original mean level 12.8 mg dL 1). The most important finding, however, was that other sources (paint) had more important impact on blood Pb than did exterior soil. There was some evidence that the biokinetics of reduction of blood Pb were slower than anticipated, with the second year of reduced exposure yielding somewhat lower blood Pb than 1 year of reduction. As noted earlier, a meta-analysis of the contribution of soil versus house dust to blood Pb of urban children has shown that house dust was considerably more important (Lanphear et al., 1998). It was suggested that the proportion of exposure from soil versus house dust in the IEUBK model needed to be changed to reflect this improved knowledge, but that has not yet occurred. 6.1.2. Bioavailability of soil/dust Pb The earliest tests were conducted with rats (Chaney et al., 1984, 1989; Stara et al., 1973) and urban dust or garden soil samples with varied levels of Pb from the Baltimore, MD, area. An ARCO Coal Company led team investigated chemical speciation of Pb in Superfund soils, and tested the bioavailability to rats and rabbits. Several groups noted that Pb in mining site soils caused less increase in blood Pb than did Pb in smelter site soils (Steele et al., 1990). Davis et al. (1993) found galena and anglesite (PbSO4) with rinding in mine waste soils at Butte, MT. The dissolution of PbSO4 was found to be inherently slow compared to the time needed for clearance of the stomach and intestine of children, helping to explain why this form of Pb caused less uptake into blood. Mineral PbS or chemical PbS both had low bioavailability to rats (Dieter et al., 1993; Fig. 4) while extremely fine PbS formed by adding sulfide to a solution of Pb isotope was somewhat more bioavailable in fasting human tests (Rabinowitz et al., 1980). Thus, part of the lower risk of mining waste PbS has to do with particle size; very fine PbS in cosmetics is apparently more dangerous than PbS in soils. Freeman et al. (1993) conducted tests of methods to assess Pb bioavailability, and found that the usual one dose, area under the curve approach was not workable for environmental levels of Pb exposure. Gavage Pb acetate was absorbed very quickly and disappeared from blood quickly, while soil Pb caused only a small increase in blood Pb over several days. So they moved to a chronic feeding approach using purified diets. It had already been shown that using ordinary rat chow greatly reduced Pb bioavailability, but using the American Institute of Nutrition (AIN) purified diets for rats promoted Pb bioavailability (Mylroie et al., 1978). They fed several levels of two soils compared to Pb acetate (Fig. 5). A number of years later Superfund researchers decided to assess RBA so that more precise decisions about soil treatment could be generated. The lead scientist in that project asserted that juvenile swine were the only animal that could be used to conduct such tests, even if rat tests were conducted to prevent the problems with rat feeding experiments

31


350

350 Data of dieter et al. (1993) Mean ± stnd error

Femur Pb, mg/kg DW

300

300 PbOAc

250

250 200

200 PbO

150

150 100

100 50

50 Pb-Ore

0

0

PbS

0

20 40 60 80 100 Diet Pb concentration (in AIN-76 diet), mg/kg DW

Figure 4 Absorption of Pb from different Pb compounds added to the AIN-76 purified diet (Dieter et al., 1993).

80 70 Bone Pb, mg/g FW

80

Female rats fed 30 days. Freeman et al., 1992. Mine waste soils. AIN-76 purified diet Mean ± SE

60 50

70 60

Ac PbO

40

ed in fe

50 40 30

30 20 10

810

ppm

oil Pb s

20

b soil

mP 8 pp

10

390

0

0

−10

−10 0

25

50

75 100 125 150 175 Measured feed Pb, mg/g

200

225

Figure 5 Effect of soil dose and mining waste soil Pb concentration on Pb in bone of rats fed the test soils in AIN diets for 35 days (Freeman et al., 1992).

(Weis and Lavelle, 1991). After years of research, that team kept asserting that it might be appropriate to use RBA measured using the swine bioassay to adjust soil Pb clean-up levels, but that the IEUBK computer program had to be used and that treatment of soils could not achieve reduction in risk. Over this decade extensive research showed that the bioavailability of soil

32


Pb could be reduced by treatment with biosolids or with phosphate sources (Brown et al., 2003; Ryan et al., 2004). A summary of that progress is reported by Ryan et al. (2004) regarding a large multi-year field test of soil Pb remediation at Joplin, MO. Eventually, the EPA Region-8 team (Casteel et al., 1996, 1997; Drexler and Brattin, 2007; USEPA, 2007a) worked to develop an independent bioaccessibility test. How to conduct relative soil Pb bioavailability tests has been debated for many years. Some felt that one had to follow the usual toxicological approach and use fasting animals. However, comparison feeding tests of Pb on fasting or to persons consuming small meals showed that for some time before and after consumption of a meal, a remarkable reduction in Pb absorption occurred in humans ( James et al., 1985). It was common for water soluble Pb to be absorbed between 60% and 80% on fasting, but only 1–5% with food (Heard and Chamberlain, 1982; Heard et al., 1983; James et al., 1985). Many have noted this critical effect and wondered how animal testing should be conducted. Is all soil/dust ingested by children ingested between meals such that only the fasting condition is representative? The juvenile swine test compromised and has fed soil mixed with a ball of the purified ‘‘dough’’ AIN feed used in the test; one can only gavage a young pig so many times before it become too difficult to continue gavage dosing. Thus, one part is fed after long fasting, and the second part is fed after a short fast. So the actual test is a hybrid or confounded method (Casteel et al., 1997, 2006; USEPA, 2007a). One issue seldom discussed is the need for testing soils with high Pb concentration in order to make significant measurements in animal tissues during bioavailability evaluation. Nearly, all of the soils fed were on the order of 10–50 times higher in Pb than the current limit for Pb in bare soils of homes under HUD rules (400 mg kg 1 for bare soil). It is possible that the soil and diet factors which interact with soil Pb bioavailability are different for such highly contaminated test soil materials than in soils to which children are commonly exposed. Ultimately, human tests of soil Pb bioavailability were conducted and these should be the ‘‘Gold Standard’’ for soil Pb bioavailability. Maddaloni et al. (1998) fed human volunteers soil on fasting or with a meal using Pb stable isotopes (206Pb/207Pb ratio) to measure the absolute Pb absorption from the test dose. By selecting subjects with quite different isotope ratios than the soil to be tested, a very sensitive assay was constructed. The soil from Bunker Hill, ID, contained 2240 mg Pb kg 1 whole soil, and 2924 mg Pb kg 1 in the 8.5). Surface complexes of arsenate on iron and aluminum oxides, examined using both infrared

140 Table 1

Retention maxima for arsenic on various solids common to soils and sediments derived from adsorption isotherms at fixed pH

Adsorbent

Al oxides Gibbsite Amorphous Al hydroxide

Activated alumina Bauxite Aluminosilicates Montmorillonite Kaolinite Fe (hydr)oxides Hydrous ferric oxide Goethite Magnetite

As(V) (mmol kg 1)

pH

As(III) (mmol kg 1)

pH

– – – – – – –

References

Approximate Hingston et al. (1971)

35 4 – 15 9 – 1500 4 – 600 9 – 1600 5 – 1200 7 – 500 9 – Twofold higher for As(V) than As(III) 67 6–7 14 52 6–7 16

6.5–8.5 6.5–8.5

8 7

5 5

3 1

5 5

Anderson et al. (1976) – – Gosh and Yuan (1987) Gosh and Yuan (1987) Gosh and Yuan (1987) Using landfill leachate Frost and Griffin (1977) Frost and Griffin (1977)

3514 173

4 4

2675 173 332

8 8 8

Dixit and Hering (2003) Dixit and Hering (2003) Dixit and Hering (2003)

Hingston et al. (1971)

Two-line ferrihydrite

a

Two-line ferrihydrite on quartz sand Others Birnessite (d-MnO2) Pyrolusite Cryptomelane Calcite (CaCO3)

2000 1500 483

4.6 9.2 7.1

100 10 25 NDa

6.5 6.5

Activated carbon Humic acids

10 90–110

3–4 5.5

Not detected.

6000 6000 1206

4.6 9.2 7.1

Raven et al. (1998) Herbel and Fendorf (2006)

Lenoble et al. (2004) Thanabalsingam and Pickering (1986b) ND

5.5

Oscarson et al. (1983) Goldberg and Glaubig (1988) Gupta and Chen (1978) Thanabalsingam and Pickering (1986a)

141

142

Scott Fendorf and Benjamin D. Kocar

(Lumsdon et al., 1984; Sun and Doner, 1996) and extended X-ray absorption fine structure (EXAFS) spectroscopy (Arai et al., 2001; Fendorf et al., 1997; Manceau, 1995; Sherman and Randall, 2003; Waychunas et al., 1993), are dominated by bidentate, binuclear (double-corner sharing) moieties. EXAFS spectroscopy is nearly blind to outer-sphere complexes when inner-sphere moieties are present, but the results obtained from this technique are consistent with infrared studies of phosphate on iron (hydr)oxides (Arai and Sparks, 2001; Parfitt et al., 1975)—a factor supporting the analogous strong retention of phosphate. However, recent X-ray scatter studies (Catalano et al., 2008) illustrate that although arsenate is strongly retained by Fe (hydr)oxides, a portion of the anion is retained as an outer-sphere complex on both hematite and corundum surfaces. Arsenic desorption from ferrihydrite, goethite, and hematite under advective flow is consistent with a portion of outer-sphere complexes that are more labile (more rapid desorption) than the large fraction of inner-sphere complexes (Tufano and Fendorf, 2008). Aluminum hydroxides and aluminosilicate clay minerals may also retain appreciable concentrations of arsenate, and they exhibit a strong preference for arsenate relative to arsenite (Tables 1 and 2) (Manning and Goldberg, 1997a,b; Smith et al., 1998; Xu et al., 1988). Similarly, Mn oxides may impart a strong influence on arsenic binding. Reaction of arsenite solutions with Mn oxides such as birnessite results in extensive and rapid uptake (Oscarson et al., 1981). However, arsenic is retained as arsenate surface complexes (Manning et al., 2002) owing to arsenic oxidation by Mn (III/IV) (Driehaus et al., 1995; Manning et al., 2002; Oscarson et al., 1981). Arsenic may also bind to organic matter in soils and sediments (Grafe et al., 2001, 2002; Ko et al., 2004; Redman et al., 2002; Thanabalsingam and Pickering, 1986a), with As(V) and As(III) having maximum adsorption on humic acids at pH 5.5 and 8, respectively (Thanabalsingam and Pickering, 1986a). Arsenic(V) adsorbs onto solid-phase humic acids more extensively than As(III), with amine (NH2) groups suspected as the primary functional group responsible for arsenic retention (Thanabalsingam and Pickering, 1986a). Arsenic adsorption by humic substances is also enhanced by cation addition, particularly Fe, Al, and Mn, whereby the cations act as bridging complexes for arsenate on humic acids (Lin et al., 2004). Nevertheless, organic matter tends to be poorly correlated with total As in comparison to Fe, Al, or P (Chen et al., 2002), suggesting that its contribution to arsenic retention in soils and sediments is often limited. In contrast to arsenate, arsenite exhibits a limited affinity for most soil minerals with the exception of iron (hydr)oxides, for which it exhibits a high degree of retention (see Table 1). Ferric (hydr)oxides and magnetite, in fact, adsorb arsenite more extensively than arsenate at all but acidic conditions (Dixit and Hering, 2003; Raven et al., 1998), as illustrated for hydrous ferric hydroxide (Fig. 1). Similar to arsenate, arsenite also forms a bidentate,

Table 2 Adsorption envelopes for As(V) and As(III) on various soil and sedimentary solids

Adsorbent

As(V)aqa; As(V)adsb

Al oxides Amorphous Al hydroxide 1600; 1700 (4.5) 133; 900 (4.5) 20; 20 (2–10) Activated alumina 53.4; 26.5 (3–7) Bauxite 53.4; 26.5 (3–7) Aluminosilicates Montmorillonite 20; 0.41 (5, 12.5) 20; 0.35 (5) Kaolinite 20; 0.5 (5) Illite 20; 0.5 (4–6) Fe (hydr)oxides Hydrous ferric oxide 20; 40 (2–9) 100; 2100 (4)

pH c maximum

4–7 4–7 2–10 3–7 3–7

As(III)aq a; As(III)adsb

pH c maximum

References

Anderson et al. (1976) 20; 16 (8.5) 26; 11 (8.2) 16; 9 (8.5)

7–9.5 3–10 3–10

Goldberg (2002) Gupta and Chen (1978) Gupta and Chen (1978) Goldberg and Glaubig (1988) Goldberg (2002) Goldberg (2002) Goldberg (2002)

5–7, >11 5–7 3–9 3–7

20; 0.4 (3) 20; 0.25 (8–10) 20; 0.22 (8–9)

3–4 7–11 7–10

2–10 Co > Ni > Cd although for Fe oxides Pb > Cu has been reported and sometimes also Ni > Co ( Jackson, 1998; Sparks, 2003). In addition, carbonate, phosphate, and sulphite present in biosolids can form sparingly soluble solid phases with many metals and thus account for a substantial

190

R. J. Haynes et al.

portion of some metals present biosolids (Karapanagiotis et al., 1991). For example, during anaerobic digestion, low solubility Cu and Zn sulfides characteristically form (Nagoshi et al., 2005). The organic component also has the ability to bind to heavy metals. The heterogeneous nature of humic substances and the large number of functional groups present means that binding of metals can be regarded as occurring at a large number of reactive sites with binding affinities that range from weak forces of attraction (ionic) to stable coordinate linkages (McBride, 2000; Sparks, 2003). Indeed, mechanisms involved in metal binding to organic matter are complex and probably involve simultaneous chelation, complex formation, adsorption, and coprecipitation (Stevenson and Vance, 1989). Because of the many variables involved, there are many inconsistencies in reported selectivity orders of metals with organic matter. A generalized order is Cr3þ > Pb2þ ¼ Hg2þ > Cu2þ > Cd2þ > Zn2þ ¼ Co2þ > Ni2þ ( Jackson, 1998; Jin et al., 1996; Stevenson, 1994). As noted previously, there is often a flush of organic matter decomposition following application, and this is followed by a slow decomposition phase. It has been suggested that heavy metals bound to biosolids organic matter could be released to soil solution during decomposition and as a result metal bioavailability would increase over time (Hooda and Alloway, 1994; McBride, 1995). In fact, it is often observed that heavy metal availability is greatest immediately (the first few months) following biosolids additions and this is followed by a reduction in availability (as estimated by metal extractability and/or plant uptake) as well as a reduction in organic matter content (Bidwell and Dowdy, 1987; Hseu, 2006; Logan et al., 1997; McBride et al., 1999; Walter et al., 2002). Nonetheless, the initial high availability may well be partially due to the rapid decomposition of biosolids organic matter and the consequent release of metals. Evidently, the metals released from decomposing organic matter are rapidly readsorbed by inorganic and/or organic components in the soil/biosolids. Biosolids pH will have a substantial controlling influence on the availability of metals following land application. In general, most heavy metal cations become increasingly immobile at high pH. This is because both their adsorption onto reactive oxide surfaces and precipitation reactions are favored at high pH (Sparks, 2003). As noted in Section 4.3, since the initial pH of biosolids is typically in the range of 6–8, their application will have a liming effect on acid soils thus raising their pH (Kidd et al., 2007) and tending to reduce metal availability. The mineralizable N content of biosolids is, however, an important property in relation to their effects on soil pH. During ammonification of organic N to NHþ 4 –N, one OH ion is released per unit of N while during þ nitrification of NH4 –N to NO3 –N, two Hþ ions are released. The overall process of conversion of organic biosolids-N to NO3–N is therefore acidifying. Thus, Hooda and Alloway (1994) observed a progressive

Inorganic and Organic Constituents and Contaminants of Biosolids

191

decrease in soil pH following biosolids application to soil which was accompanied by an accumulation of soil NO 3 –N and an increase in uptake of Cd, Ni, Pb, and Zn by ryegrass growing in the soil. Such an increase in metal bioavailability accompanying acidification induced by nitrification of biosolids-derived N has also been observed by others (De Haan, 1975; Hooda and Alloway, 1993). It is therefore important to monitor pH and apply lime, if necessary, to maintain a relatively high pH (e.g., 6.5) following biosolids application. As noted in Section 4.3, the high EC of biosolids may result in an increase in soluble salts in soil solution. High soluble salts will tend to reduce soil solution pH (by exchange between cations in soil solution and Hþ and Al3þ on soil cation exchange sites) thus increasing the solubility of heavy metal cations. In addition, high concentrations of solution Cl can increase mobilization, availability, and plant uptake of Cd through the formation of Cd–chloro complexes (Weggler-Beaton et al., 2000). 5.3.3. Effects of soil properties on availability Soil properties such as pH, redox potential, EC, clay, hydrous oxide, and organic matter content will also influence heavy metal availability. The most widely recognized factor is soil pH. With the exception of As and Se, heavy metal retention by soils increases with increasing pH (McBride, 1994). As noted above, with an increase in pH, the charge on the variable charge adsorption surfaces (e.g., Fe, Al, and Mn hydrous oxides) becomes increasingly negative thus favoring metal cation adsorption and the high pH also favors surface precipitation of the metals onto the surfaces (Bradl, 2004; McBride, 2000). In general, the more mobile metals such as Ni, Cd, and Zn are more sensitive to increasing pH than other metals such as Pb and Cu that are more strongly complexed with soil organic colloids (Smith, 1996). Manipulation of soil pH has been found to be the most effective way of controlling heavy metal bioavailability in biosolids-treated soils (Alloway and Jackson, 1991). Indeed, a large number of workers have shown that the bioavailability of metals to plants in biosolids-amended soils decreases as pH is raised either by liming or applying lime-stabilized sludges (Basta and Sloan, 1999; Milner and Barker, 1989; Oliver et al., 1998). Liming a range of biosolids-treated soils to pH 7 was shown by Jackson and Alloway (1991) to reduce Cd content of lettuce by an average of 41% and cabbage by 43%. Redox potential is often considered an important factor although both increases and decreases in heavy metal solubility have been recorded following waterlogging and the onset of anaerobic soil conditions (Charlatchka and Cambier, 2000; Chuan et al., 1996; Grybos et al., 2007; Kashem and Singh, 2001a,b; Xiong and Lu, 1992). This is because a number of different processes occur following the onset of anaerobiosis and these often interact to affect metal solubility. In freely-drained soils,

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Fe and Mn occur in their high oxidation states as oxides and hydrous oxides. However, as soils become anaerobic, due to waterlogging, the redox potential decreases and oxide minerals begin to dissolve as soluble Mn2þ and Fe2þ forms (Stum, 1992; Stum and Sulzberger, 1992). This can not only result in an increase in the solubility of Mn and Fe but also of other metals (e.g., Zn, Cu, Co) which were previously adsorbed to, or occluded by, these oxides (Chuan et al., 1996; Grybos et al., 2007). When soils become anaerobic the pH tends to converge to neutrality irrespective of initial pH, whether acidic or alkaline (McBride, 1994). For acidic soils this increase in pH can result in release of organic matter into soil solution and metals bound to the organic molecules are also thought to be released (Grybos et al., 2007). This also tends to increase metal solubility. Nonetheless, the increase in pH up to about 7, favors adsorption/surface precipitation of metal cations thus favoring removal of metals from solution (Kashem and Singh, 2001a). In addition, at low redox potential sulfate ions are reduced to the sulfide form which may form complexes with metals such as Cd, Zn, and Ni (Hesterberg, 1998; Van Den Berg et al., 1998). Most metal sulfides are insoluble even under acidic conditions and so this process also tends to reduce soluble metal concentrations. Oxidation state of the contaminant itself also affects solubility. For example, selenite [Se(IV)] is much more strongly adsorbed to soil colloid surfaces than selenate [Se(VI)] and the presence of selenite is favored under reducing conditions (Martinez et al., 2006; Neal and Sposito, 1989). Se will therefore be less plant available under reducing conditions. Furthermore, under strongly reducing conditions Se may form elemental Se and metal selenides (e.g., FeSe) both of which are insoluble (Elrashidi et al., 1987; Masschelyen et al., 1991). Under oxidizing conditions both arsenate [As(V)] and arsenite [As(III)] are present while under reducing conditions As is present mainly as As(III) (O’Neill, 1995). Compared to other As species, As(III) exhibits the greatest mobility and plant availability because of its presence as the neutral species H3AsO3 (Ascar et al., 2008; Marin et al., 1993). Nonetheless, strongly reducing conditions in biosolids-amended soils can lead to precipitation of As as As2S3 (Carbonell-Barrachina et al., 1999). The ability of soils to adsorb and sequester metals is also an important factor. This is dependent on their content of inorganic (clay and Fe, Mn and Al hydrous oxide content) and organic (soil humic material) binding agents. For example, sandy soils with low oxide content and low organic matter have low sorption capacities and will have greater metal availabilities than loamy or clayey soils containing greater amounts of sorbents (e.g., clays, oxides, and organic matter) provided the soils have similar pH values (Basta et al., 2005). Hue et al. (1988) applied increasing rates of biosolids to three different soils, a limed volcanic ash-derived Andept, an alkaline Vertisol, and a limed manganiferous Oxisol. DTPA-extractable soil metal levels, lettuce growth, and tissue metal concentrations were measured.

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The Andept had the highest metal adsorption capacity and the Oxisol the lowest. As a result, lettuce Cd, Mn, Ni, and Zn concentrations were highest in the Oxisol and Mn levels reached phytotoxic levels. Hue et al. (1988) concluded that the Andept could tolerate the highest biosolids loading rate and the Oxisol the lowest. The calcite (CaCO3) content of soils can also be important. In calcareous soils, calcite represents an effective sorbent for metal ions. The initial reaction is thought to be chemisorption but metals with an ionic radius similar to that of Ca (Cd 2þ, Mn 2þ, Fe 2þ) can also readily enter the calcite structure and form coprecipitates (Gomez de Rio et al., 2004; McBride, 2000). 5.3.4. Metal availability over time The long-term (>10 years) bioavailability of heavy metals in biosolidsamended soils is of great importance in relation to environmental effects of land application of biosolids. As noted previously (Section 5.3.2), following a one-off application of biosolids the extractability of metals generally declines over time (Hseu, 2006; Sukkariyah et al., 2005a; Walter et al., 2002). Sukkariyah et al. (2005a), for example, showed DTPA-extractable Cu and Zn levels progressively decreased following one-time applications of biosolids at rates ranging from 42 to 210 Mg ha 1 (Table 3). Seventeen years after application, extractable concentrations of Cu and Zn had decreased by 58% and 42%, respectively. The decrease is attributable to metals reverting to more recalcitrant forms in the soil such as occlusion in Fe oxides or chemisorption to surfaces. Despite the initial decrease in extractability, concentrations of extractable heavy metals in biosolids-amended soils can remain elevated above Table 3 Long-term effect of biosolids application on DTPA-extractable Cu and Zn DTPA-extractable Zn mg kg 1

DTPA-extractable Cu mg kg 1

a

Biosolids rates Mg ha 1

1984

1995

2001

1984

1995

2001

0 42 84 126 168 210

1.4f a 24.9e 53.0d 73.4c 119.9b 129.4a

3.7f 23.1e 44.3d 64.8c 78.7b 92.8a

3.2f 12.6e 25.4d 33.7c 43.3b 53.6a

1.6f 19.2e 38.9d 52.4c 73.2b 78.2a

2.8f 17.2e 33.3d 49.6c 59.5b 69.9a

2.7f 9.1e 19.8d 27.9c 35.5b 49.7a

Values within columns followed by different letters are significantly different at the 0.05 probability level. From Sukkariyah et al. (2005a); copyright American Society of Agronomy.

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those of control for many decades after applications have ceased (Alloway and Jackson, 1991; Basta et al., 2005; McBride, 1995; McGrath, 1987). Results from a long-term market garden experiment at Woburn (UK) serve to illustrate this point. Sludge was applied in the 1940s until the 1960s and CaCl2-extractable Cd changed little from 1950 until the early 1980s remaining significantly higher than the control soils over the entire interval monitored (McGrath and Cegarra, 1992). Similarly, EDTA-extractable Cu, Pb, Zn, Ni, and Cr changed little following termination of biosolids application and treated soils maintained a much greater proportion of metal in EDTA-extractable form than the control. Such results occurred despite there being a significant loss of biosolids organic matter over the period indicating that heavy metals released from the decomposing organic matter were rapidly adsorbed by inorganic components of biosolids/soil and/or native soil organic matter. Certainly, biosolids-derived heavy metals are strongly sorbed to soil components making them characteristically immobile in soils. Indeed, the vast bulk of the added metals remain in the topsoil in the layer of incorporation and there is a marked reduction in concentration with depth (Alloway and Jackson, 1991; Brown et al., 1997; Chang et al., 1983; Sloan et al., 1997; Sukkariyah et al., 2005b). 5.3.5. Heavy metal mobility and leaching The results of Sukkariyah et al. (2005b) serve to illustrate the immobility of biosolids-borne heavy metals in soil. They found that more than 85% of total applied Cu and Zn was still in the layer of incorporation (0–15 cm) 17 years after a one-time biosolids application. Results for Mehlich I-extractable Cu and Zn at that site are shown in Fig. 3. It is evident that extractable Cu and Zn are concentrated in the 0–15 cm layer but there is some indication of a small amount of movement down into the 15–20 cm layer. Mass balances calculated for several long-term experiments do suggest some losses of heavy metals from the topsoil (McBride, 1995). Lateral movement in the soil due to tillage (McGrath and Lane, 1989) or physical mixing with the lower soil layer by plowing (Sloan et al., 1998) can be responsible for a significant part of the losses from the original amended soil layer. Nevertheless, mass balances calculated for sites where little or no tillage has been performed have shown less than 100% recovery (McBride et al., 1999). Increased extractable heavy metal levels (e.g., for Cu, Zn, Ni, Pb) at depths of 20–150 cm below the level of incorporation have been noted in field experiments (Barbarick et al., 1998; Baveye et al., 1999; Bell et al., 1991; Keller et al., 2002; Schaecke et al., 2002). Leachate sampling below field plots and/or undisturbed monolith lysimeters receiving biosolids has also revealed elevated metal concentrations (Keller et al., 2002; Lamy et al., 1993; McBride et al., 1997, 1999; Richards et al., 1998; Sidle and Kardos, 1977). In addition, column leaching studies have shown that heavy metals can leach through many tens of cm of soil (Al-Wabel et al., 2002;

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0

10

Concentration (mg kg−1) 20 30 40 50 60

70

0–15 Cu 15–20

20–25

25–30

30–35 //

//

80–85

210 Mg ha−1 126 Mg ha−1 Control

Depth, cm

85–90 0

20

40

60

80

0–15 Zn 15–20

20–25

25–30

30–35 //

//

80–85

210 Mg ha−1 126 Mg ha−1 Control

85–90

Figure 3 Distribution of Mehlich-I extractable Cu and Zn with soil depth 17 years after biosolids application. From Sukkariyah et al. (2005a,b); copyright American Society of Agronomy.

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Antoniadis and Alloway, 2002; Ashworth and Alloway, 2004; Parakash et al., 1997; Toribio and Romanya, 2006). In most studies, the annual export of metals from the surface-mixing layer represents a small fraction (i.e., 99% of soluble Cu and Zn in leachates was present in organically complexed form. Heavy metals have, however, also been shown to be present in drainage water associated with suspended clay-sized particles (Keller et al., 2002). The metals become adsorbed to the surfaces of Fe oxide and layer silicate clays present in this leached particulate matter. Keller et al. (2002) calculated that movement of particulate matter accounted for about 20% of Cu, Zn, and Cd leaching from a biosolidsamended soil. An important factor thought to contribute to leaching of metals is preferential flow of water and dissolved metals down the soil profile in downward oriented macropores (e.g., cracks, earthworm channels, root channels) (Camobreco et al., 1996; Keller et al., 2002; Lamy et al., 1993). This water bypasses the soil matrix thus minimizing the chances that the dissolved metals will be adsorbed to soil surfaces. Preferential flow is probably the main pathway of movement of suspended particulate matter and associated metals (Keller et al., 2002). The period of greatest risk of metal leaching is soon after biosolids application. This is when soluble organic matter is present in high concentrations and when preferential flow down surface-connected macropores is most likely. Indeed, leaching losses of metals are normally greatest during this initial period (Antoniadis et al., 2007; Camobreco et al., 1996; Keller et al., 2002; Lamy et al., 1993; Maeda and Bergstrom, 2000). For this reason, it will be important to minimize water inputs (e.g., irrigation) and drainage from soils immediately following land application of biosolids. 5.3.6. Soil microbial/biochemical effects Elevated concentrations of heavy metals in soils are known to affect soil microbial populations and associated activities (Baath, 1989; Brookes, 1995; McGrath, 1994). Baath (1989) concluded that the following order of toxicity to soil microbes is most commonly found (in mg kg 1 values): Cd > Cu > Zn > Pb. However, he showed an enormous disparity between individual studies as to the exact concentrations at which metals become toxic. Giller et al. (1998) suggested that much of the variability in deriving toxic concentrations of heavy metals occurs through comparison of results from short-term laboratory incubation studies with data from long-term exposures of microbial populations to heavy metals in field experiments. This is because laboratory studies measure response to immediate acute


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toxicity (usually from one large addition of metals) whereas monitoring of long-term field experiments measures responses to long-term chronic toxicity which accumulates gradually. Stress caused by heavy metal contamination typically has two interrelated effects on soil microbial communities. The first is a loss of structural and functional diversity since toxicities can suppress and/or kill sensitive parts of the community. Nevertheless, rediversification can occur in the surviving tolerant communities (Barkay et al., 1985). The other is an increase in respiration per unit of microbial biomass (metabolic quotient; qCO2) which is thought to occur because stressed microorganisms direct a relatively larger amount of available energy into maintenance of various biochemical functions (Giller et al., 1998). Thus, in general heavy metal contamination of soils has been shown to result in a decline in microbial biomass C, an increase in metabolic quotient (Brookes, 1995; Giller et al., 1998), and shifts in bacterial community structure (Frostegard et al., 1996; Giller et al., 1998; Tom-Petersen et al., 2003). There are also often negative effects on soil enzyme activity (Belyaeva et al., 2005; Kizilkaya and Bayrakli, 2005). Enzyme reactions can be inhibited by heavy metals through a number of mechanisms including by (i) complexing with the substrate, (ii) combining with the protein-active groups of the enzymes, or (iii) reacting with the enzyme–substrate complex (Dick, 1997). In the case of biosolids application to soils, the addition of organic material increases organic matter content and consequently the size and activity of the microbial community also tend to be stimulated (Section 3.1.2). However, if biosolids contain a high heavy metal load then metal toxicities may have an inhibitory effect on soil microbial activity. Indeed, many workers have observed an inhibitory effect in soils where biosolids high in heavy metals have been applied and these negative effects can remain for decades after application (Giller et al., 1998; Stoven et al., 2005). Numerous short- and long-term studies have been carried out where biosolids contaminated with one or more heavy metals (or biosolids enriched with one or more heavy metals) have been applied to soils and the size and activity of the microbial community measured. Short-term incubation experiments have generally shown a reduction in microbial biomass C and N, usually an increase in metabolic quotient and a variable effect on enzyme activity (Bhattacharyya et al., 2008; Kao et al., 2006; Rost et al., 2001). Long-term (>8 years) field trials have shown similar results with a depression in microbial biomass C and microbial biomass C expressed as a percentage of organic C and an increase in metabolic quotient (Bhattacharyya et al., 2008; Chander and Brookes, 1991; Fliebßach et al., 1994; Stoven et al., 2005; Zhang et al., 2008). Zhang et al. (2008) sampled soils in fields that had been irrigated with heavy metal contaminated wastewater (polluted with Cd and to a lesser extent Zn and Cu) for

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30 years along a gradient of increasing total soil Cd content (1–4) (Table 4). Concentrations of extractable Cd, Cu, and Zn and metabolic quotient generally increased along the gradient while microbial biomass C declined (Table 4). Observed effects on soil enzyme activities have been variable with Bhattacharyya et al. (2008) observing reductions in glucosidase, urease, phosphatase, and sulphatase activities induced by high combined concentrations of Cd, Cr, Cu, and Pb, Zhang et al. (2008) finding dehydrogenase and phosphatase activities were not consistently affected by a combination of high Cd, Cu, and Zn (Table 4) and Stoven et al. (2005) finding dehydrogenase activity was decreased but that of phosphatase was unaffected by high combined concentrations of Cr, Cd, Cu, Hg, Ni, Pb, and Zn. Not only is the size and activity of the soil microbial community affected by heavy metal contamination originating from biosolids but also its composition is altered (Macdonald et al., 2007; Sandaa et al., 1999a,b). Bioluminescence-based bacterial and fungal biosensors can be used to assay the potential toxicity of water-soluble contaminants in soils and this technique was employed by Horswell et al. (2006) to determine the effects of Cu-, Ni-, and Zn-spiked biosolids on the microbial community in the litter layer of a forest soil. They found that increased Cu caused a decline in bioluminescence response of the fungal biosensor, increased Zn caused decline in response of the bacterial biosensor while increased Ni had little effect on either. In a 10-year field experiment where plots received different concentrations of biosolids spiked with a combination of Cd, Cu, Ni, and Zn, molecular techniques were used to show that significant differences, and decreased diversity, were induced in both bacterial (Sandaa et al., 1999a, 2001) and archaeal (Sandaa et al., 1999b) community structures. Using molecular techniques Macdonald et al. (2007) showed that in an 8-year study using Zn-spiked biosolids there were significant differences in microbial community structure for all groups investigated (bacteria, fungi, archaea, actinobacteria, and rhizobium/agrobacterium). Their results showed that fungi, and to a lesser extent archaea, were more negatively affected by Zn addition than was the bacterial community. Results from several long-term experiments have shown that Rhizobium leguminosarum, a N2-fixing symbiotic bacteria of white clover, is considerably more sensitive to the toxic effects of heavy metals than the host plants and that the host plant confers protection from metal stress to the rhizobium (Chaudri et al., 1993; McGrath et al., 1995). The toxic effect is due to toxicity to the free living rhizobium particularly in response to high Zn (Chaudri et al., 2008). Thus, the general effect of heavy metal contamination of soils induced by biosolids applications is a decrease in the size of the microbial community, an increase in metabolic quotient, a change in species composition, and often a decrease in activity of key enzymes involved in C, N, P, and S transformations. Such decreased enzyme activity will tend to reduce the turnover of C, N, P, and S in the soil. The potential effect of a change in

Table 4 DTPA-extractable Cd, Cu, and Zn, microbial Biomass C, metabolic quotient, and the activities of dehydrogenase and cellulose in soils on a gradient of increasing Cd loading from 30 years of irrigation with heavy metal-contaminated wastewater

Site

1 2 3 4

DTPA-extractable (mg kg 1) Cd

Cu

Zn

0.48 0.48 0.52 1.05

5.18 5.00 5.28 13.0

3.42 4.62 2.44 9.32

From Zhang et al. (2008); copyright Elsevier.

Microbial biomass C (mg kg 1)

Metabolic quotient (102 mg CO2–C h 1 mg 1)

Dehydrogenase activity (mg product kg 1 h 1)

Cellulase activity (mg product kg 1 h 1)

207 199 177 142

5.47 5.17 4.91 7.51

1.55 4.21 3.89 2.02

14.8 10.3 16.8 14.3

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species composition and loss of some microbial species is controversial. This is because there is a high degree of functional redundancy among soil microflora and functionally similar organisms can have different environmental tolerances (Nannipieri et al., 2003). Thus, as a stress is imposed on the soil, there will be a progressive extinction of sensitive species among a functional group; below a certain threshold it is speculated that there will be insufficient individuals to sustain a particular function (Giller et al., 1997). Such a situation does not appear to have occurred in biosolids-treated soils. With increasingly stringent regulations being enforced regarding the concentrations of heavy metals that can be released into sewerage systems, the negative effects of biosolids applications on soil microbial activity, induced by their heavy metal loads, is likely to decline over the ensuing years.

5.4. Plant response and metal uptake At agronomically realistic rates of biosolids application (e.g., 2–8 Mg ha 1), heavy metals do not normally represent a serious limitation to crop growth even though relatively large amounts of metals can be added to the soil; indeed as noted previously, the majority of experiments demonstrate a positive effect of biosolids application on crop growth ( Juste and Mench, 1992; Singh and Agrawal, 2008). For example, Juste and Mench (1992) reviewed a large number of field trials in the United States and Europe and concluded that phytotoxicity due to biosolids-borne metals was only rarely observed in grain crops. At one site, Cd toxicity, Ni toxicity, or the combined effects of Cd and Ni caused yield reductions in maize. At very high rates of biosolids application, metal toxicities are likely to occur. Berti and Jacobs (1996), for instance, applied large amounts of biosolids to croplands over a 9-year period (cumulative applications of 240–690 Mg ha 1) and recorded yield reductions in maize, sorghum, and soybean due to the combined phytotoxic effects of Zn and Ni. Leguminous crops are usually more sensitive to metal loadings in biosolids than nonlegumes (Giordano et al., 1975; McGrath et al., 1988). Such effects are explicable in terms of the detrimental effects of heavy metals on Rhizobium nodule function. There are a number of reasons for the low phytotoxic effect of biosolidsborne metals. These can include metal sorption by metal oxides and organic matter (present in both the biosolids and the soil), increased pH, formation of insoluble salts (e.g., with silicate, sulfate, and phosphate), and antagonistic effects of between biosolids metals (Bell et al., 1991; Emmerich et al., 1982; Juste and Mench, 1992). Despite this, heavy metal loadings are often considered the major potentially detrimental effect of land-applied biosolids on the environment. Not only do they accumulate in the soil but also phytotoxic effects on plants can occur. More concerning is the accumulation of metals in crops in edible plant parts. These aspects are considered below.


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5.4.1. Metal toxicity and tolerance Elevated concentrations of both essential (e.g., Cu, Zn) and nonessential (e.g., Cd, Pb, Hg) heavy metals in soils can inhibit plant growth. Toxicity is caused by a range of interactions at both the cellular and molecular levels (Clemens, 2006; Clemens et al., 2002; Hall, 2002; Shaw et al., 2006). It can result from binding of metals to sulfhydryl groups in proteins leading to an inhibition of activity or disruption of structure, from displacement of an essential element resulting in deficiency effects or from stimulation of the formation of free radicals and reactive oxygen species resulting in oxidative stress (Dietz et al., 1999; Hall, 2002; Van Assche and Clijsters, 1990). Metal tolerance is ubiquitous and the main components of metal homeostasis in plants involve transport, chelation, and sequestration processes (Clemens, 2001, 2006; Clemens et al., 2002; Hall, 2002). The regulated activities of these processes ensure proper delivery and distribution of metals at both the organizational and cellular levels (Clemens, 2001). Loss of one of these critical processes (e.g., synthesis of chelating agents for Cd within cells) leads to genotypes that are more sensitive (hypersensitive) than wild-type plants (Howden et al., 1995). By contrast, some plant species and genotypes can grow naturally in soils containing concentrations of metals that would be toxic to the majority of plants. These hypertolerant species possess naturally selected higher levels of tolerance. Potential mechanisms of detoxification are primarily involved in avoiding the buildup of toxic concentrations at sensitive sites within the cell (Hall, 2002) thus preventing the damaging effects outlined above. Some plants not only tolerate high concentrations of metals but also hyperaccumulate them. Hyperaccumulator is a term used to describe plants capable of accumulating more than 1000 mg g 1 Ni, Co, Cu, Pb, for Cd 100 and Zn 10,000 mg g 1 (Love and Babu, 2006). Metal hyperaccumulation is a rare phenomenon limited to about 400 different species belonging to a wide range of taxa (Baker and Brooks, 1989). Hyperaccumulation is mainly observed with Ni, Zn, Co, and Se but there are also reports of it for Pb, Cd, and As (Baker and Brooks, 1989). The existence of hyperaccumulator plants is the basis for the phytoremediation technique in which plants are used to absorb metals and hence remove them from contaminated soils (Chaney et al., 1997). Possible strategies of heavy metal tolerance in plants are diverse and can be extracellular, at the plasma membrane, or intracellular. Extracellular mechanisms include the effects of ectomycorrhizae ( Jentsche and Godbold, 2000) and arbuscular mycorrhizae (Kaldorf et al., 1999; Leyval et al., 1997) in restricting metal movement into the plant host roots. This has been attributed to absorption of metals by the hyphal sheath, chelation by fungal exudates, adsorption onto external mycelium, and reduced access to the plant root cytoplasm ( Jentsche and Godbold, 2000; Leyval et al., 1997). Root exudates (e.g., organic acids such as citrate, malate, and oxalate) have

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also been suggested as agents for detoxification through chelation of metals in the rhizosphere (Dong et al., 2007). In solution culture experiments, addition of a range of organic compounds, such as organic acids, amino acids, and peptides, to the growing medium can alleviate heavy metal toxicities (Hall, 2002) and the Ni chelating ability of citrate and histidine in root exudates has been associated with its nonaccumulation in species of Thlaspi (Salt et al., 2000). Another possible mechanism is the accumulation of heavy metals in root cell walls. Indeed, in many plants that accumulate heavy metals, 60–80% of metals (e.g., Cd, Cu, Zn, Pb, Co, Ni) accumulate in roots and root cell walls are the main region of accumulation (Liu et al., 2007a,b; Nishizono et al., 1987; Sousa et al., 2008; Zornoza et al., 2002). In some cases, this accumulation in the cell wall may reflect active efflux of metals from root cells through the plasma membrane (see below). Furthermore, for some plants (e.g., lettuce) that accumulate metals (e.g., Cd) in leaves, the Cd accumulates principally in the cell walls (Ramos et al., 2002) suggesting it has been deposited there by active efflux across leaf cell plasma membranes. There are three suggested mechanisms at the plasma membrane. Firstly, plasma membrane function may be rapidly affected by high concentrations of metals with increased leakage of solutes (e.g., K) occurring (Hall, 2002). Tolerance could involve protection of plasma membranes against metal damage or improved repair mechanisms (Hall, 2002). The cell membrane could also play an important role in reducing entrance of metals into cells (Clemens, 2006). However, this strategy is thought to be of minor significance since many essential heavy metals must be taken up from the soil for various metabolic functions and the chemical properties of nonessential metals are similar to those of essential ones and their uptake is thought to occur through the same processes/mechanisms (Shaw et al., 2006). An alternative strategy is active efflux of toxic metal ions. In bacteria, most resistance systems are, in fact, based on energy-dependent efflux of toxic ions either by ATPases or chemostatic cation/proton antiporters (Silver, 1996). It seems likely that such a mechanism also occurs in higher plants (Hall, 2002). It is generally accepted that the principal mechanism of detoxification and tolerance is achieved by sequestration of metals inside cells (Shaw et al., 2006). Chelation of metals in the cytosol by organic ligands is an important intracellular mechanism of detoxification and tolerance. Chelates bind with metals and buffer cytosolic ionic metal concentrations. Carboxylic acids (e.g., citrate, malate, oxalate) and amino acids such as histidine are potential ligands (Clemens, 2001) and metallothioneins (low molecular weight, cysteine-rich proteins) may also play a role (Cobbett and Goldsbrough, 2002). Nonetheless, the most studied ligands are phytochelatins which are metalbinding peptides synthesized in plants in response to exposure to metal ions (Cobbett and Goldsbrough, 2002; Rauser, 1999). Hg, Cd, As, and Cu show


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the greatest tendency to induce phytochelatins and Zn, Pb, Ni, and Se have a lesser effect (Shaw et al., 2006). For Cd, the phytochelation pathway consists of two parts. Firstly, metal-activated synthesis occurs and metals in the cytoplasm are complexed. The second stage is active efflux of the metal– phytochelatin complex across the tonoplast and sequestration in the vacuole. The vacuole is known to be the site of accumulation of metals such as Zn and Cd (Ernst et al., 1992). As already noted, active efflux of metals across the plasma membrane with their accumulation in cell walls may be another important mechanism. It is evident that the general strategy of plants is to avoid toxicity by minimizing the buildup of excess metals in the cytosol. This is achieved through the use of a variety of mechanisms. It is likely that a specific mechanism, or combination of mechanisms, is employed for a specific metal in particular plant species. In situations where soils have become highly contaminated with heavy metals, the use of tolerant species/cultivars that employ these mechanisms can be an important consideration. However, since heavy metal toxicities in plants are rarely encountered in biosolids-amended soils, this aspect is of minor importance to land application of biosolids. Accumulation of heavy metals in plants and plant parts can, nevertheless, be of concern. 5.4.2. Metal accumulation in plants One of the major concerns regarding land application of biosolids is that heavy metals may accumulate in plants and subsequently enter the food chain and/or have toxic effects on humans or grazing animals ingesting them. Indeed, the main route of entry of metals into human food chain is their accumulation in edible portions of crop plants and this can pose a potential threat to human health (McLaughlin et al., 1999). Chaney (1980) classified metals into four groups when considering their potential health risks. Group 1 (Ag, Cr, Sn, Ti, Y, and Zr) were considered to pose little risk because their low solubility in soils results in them not being taken up by plants to any great extent. Group 2 (As, Hg, and Pb) are also strongly adsorbed to soil surfaces and while they may be taken up by plants they are not readily translocated to edible portions. They therefore pose little risk to human health. Group 3 (Cu, Mn, Ni and Zn) are accumulated in plants but are phytotoxic at concentrations that pose little risk to human health. Members of Group 4 (Cd, Co, and Se) can pose human health risks at plant tissue concentrations that are not phytotoxic. In general, metals that have most commonly given rise to human health concerns in relation to food safety are Cd, Hg, Pb, As, and Se (Reilly, 1991). McLaughlin et al. (1999) considered that Cd and Se are of greatest concern in relation to terrestrial food chain contamination. In particular, excessive human intake of Cd is of concern since it accumulates in the body and impairment of kidney function is the main adverse effect.

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Concentrations of extractable heavy metals in soils are likely to give the best estimate of metal phytoavailability. Even so, at a broad level, metal concentrations in plants grown on biosolids-treated soils are a function of the annual biosolids loading rate ( Juste and Mench, 1992). Cumulative metal input to the soil is also a major factor determining plant tissue metal concentrations (Chang et al., 1997; Soon et al., 1980). Nonetheless, although plant metal concentrations generally increase with increasing biosolids rates, concentrations in plant tissues often exhibit a plateau response at high loadings (Basta et al., 2005). Indeed, a number of studies have showed that metal uptake reaches a maximum with increasing biosolids application for wheat, maize, and a range of vegetables (Barbarick et al., 1995; Brown et al., 1998; Logan et al., 1997, Sukkariyah et al., 2005a). Although a number of explanations have been forwarded for this (Basta et al., 2005, McBride, 1995) the most likely explanation is that metal availability in biosolids-treated soils shows a plateau at high loadings corresponding to metal availability in the biosolids (Basta et al., 2005; Corey et al., 1987). Plant tissue metal concentrations will, however, not only be related to biosolids application rate but will also be both metal and plant specific. That is, the mobility of individual metals in plants differs and the ability of individual plant species and cultivars to absorb and translocate metals also differs. The transfer coefficient (TC) (the concentration of metal in the plant to that in the soil) gives an indication of its mobility. From a number of studies, Antoniadis et al. (2006) calculated TC values of Cr 0.0005, Pb 0.02, Ni 0.06, Cu 0.21, Cd 0.94, and Zn 1.05. Similarly, Alloway (1995) also noted that TC values were highest for Cd and Zn (1–10) and least for Cr and Pb (0.01–0.10). In a review of a number of long-term (>10 years) biosolids trials, Juste and Mench (1992) concluded that Cd, Ni, and Pb were the most likely to accumulate in plants while for Cr and Pb, plant uptake was insignificant. Generally, metals are present in highest concentrations in roots and lowest concentrations in seeds (Love and Babu, 2006). However, Cd can accumulate in leaves so on Cd-contaminated soils, leafy vegetables can be as much, or more, of a risk than seed or root crops (Alloway and Jackson, 1991). The propensity for plants to accumulate and translocate metals to edible and harvestable parts depends greatly on plant species as well as agronomic, climatic, and soil factors. Different plant species and cultivars are known to accumulate different quantities of metals when grown in the same contaminated soil. Antoniadis et al. (2006) broadly classified crop species in relation to metal accumulation. Crops with a high plant uptake included lettuce, spinach, celery, kale, ryegrass, sugarbeet, and turnip while the low uptake group included potato, maize, peas, leek, onion, tomato, and berry fruit. Nonetheless, plant species differ in their capacity to accumulate individual heavy metals. For example, on biosolids-amended soils Davis


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and Carlton-Smith (1980) found the trend for metal accumulation was tobacco > lettuce > spinach > celery > cabbage for Cd; kale > ryegrass > celery for Pb; sugarbeet > some varieties of barley for Cu and sugarbeet > mangold > turnip for Zn. For Cd, Alloway et al. (1990) showed Cd accumulation followed the order: lettuce > cabbage > radish > carrots. Maize cultivars have been shown to differ considerably in their uptake of both Cd and Zn (Hinesly et al., 1982; Logan and Miller, 1985). For root crops grown in biosolids-amended soils, it may be important to wash of all adhering soil (or peel them) to avoid soil contamination with heavy metals. Indeed, where crops are being grown on biosolids-amended soils, routine monitoring of heavy metal concentrations in edible plant parts is an important consideration. Many countries have set limits for metal concentrations in foods and it is important that these are not exceeded.

5.5. Ingestion by animals Pasture is an attractive option for land application of biosolids since there is greater accessibility for a greater proportion of the year compared to arable land. Nevertheless, a major concern is transfer to, and accumulation of, heavy metals in the grazing animal and the human food chain (Hill, 2005; Hillman et al., 2003). The three main pathways by which this may occur are (i) transport of metals from soil to plants and then ingestion by animals (biosolids–soil–plant–animal), (ii) direct contamination of plants subsequently fed to animals (biosolidsplant–animal), and (iii) ingestion of contaminated soil by grazing animals (biosolids–soil–animal) (Fries, 1996). Accumulation of metals into foliage of grasses is generally low since metals usually accumulate in the roots (Love and Babu, 2006). Thus, pathway (i) is likely to be of minor importance. However, where surface application of biosolids is practiced, adhesion of biosolids-derived metals to pasture herbage occurs (Aitken, 1997; Klessa and Desira-Buttegieg, 1992). The greatest intake will occur when biosolids are applied directly to established pasture and animals have immediate access. Intake will be reduced if access to pasture is delayed so there is time for biosolids to be washed off leaves, foliar concentrations are reduced by dilution through plant growth and the biosolids can move onto/into the soil surface (Aitken, 1997; Hillman et al., 2004). Herbage may also become contaminated through rain splash or from deposition of dust (Aitken, 1997). There are waiting periods imposed in most countries/states/regions following surface application of biosolids to pastures before land can be grazed again. These are generally formulated to minimize the risk of exposure of grazing animals to pathogens rather than to chemical contaminants and can range from as short as 3 weeks in the United Kingdom, Spain, and the Netherlands to 1 year in Denmark (Hill, 2005).

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When pasture is the main diet of grazing animals, the chance for animal ingestion of biosolids-borne metals is greatest since soil ingestion becomes an additional important pathway. Research has shown that when pasture is the sole animal feed source, soil ingestion is generally inversely related to the availability of forage (Beyer and Fries, 2003). Lowest soil ingestion rates (1–2% of dry matter intake) occur in spring when grass growth is greatest while when forage is sparse (autumn and winter) soil intake can be as great as 18%. For sheep that graze close to the ground soil ingestion rates of as high as 30% have been reported on occasions (Abrahams and Steigmajer, 2003). In addition, in winter, poaching, or pugging of pasture can also occur (Drewry et al., 2008). This is when the hooves of grazing sheep and cattle trample the foliage into muddy soil thus causing soil contamination of forage. For grazing sheep and cows (fed no additional forage) average annual soil ingestion commonly amounts to 4.5–5.1% of dry matter intake (Beyer and Fries, 2003). In general, soil/biosolids ingestion is the main pathway for transfer of metals from pasture to the grazing animal (Beresford and Howard, 1991; Hillman et al., 2003; O’Riordan et al., 1994; Rafferty et al., 1994). There is differential accumulation of individual metals in tissues of grazing ruminants with Cd and Pb accumulating in the liver and kidney to a greater extent than muscle and fat tissues (Hillman et al., 2003). Cu tends to accumulate in the liver while Zn and Fe have no clear patterns of accumulation. The percentage of metals retained is generally low (Baxter et al., 1982; Johnson et al., 1981). For example, Johnson et al. (1981) fed steers a diet containing 11.5% biosolids for 106 days. Retention of ingested metals averaged 0.09%, 0.06%, and 0.30% for Cd, Hg, and Pb, respectively, and no retention of Cu and Zn was detected. Nonetheless, this low percentage retention increased tissue Cd, Hg, and Pb concentrations in liver and kidney by 5–20-fold. Overall, the risk to the human food chain from biosolids-derived heavy metals via grazing animals is considered low (Hillman et al., 2003).

6. Organic Contaminants In the last 50-years production of synthetic organic chemicals for industrial and domestic uses has increased enormously. For example, between 1950 and 1970, production increased from 7 million tons to 63 million tons (Rogers, 1996). As a consequence, the occurrence and concentration of organic contaminants in effluents, sewage, and biosolids has also increased. The presence and level of organic contaminants in biosolids depends greatly on the quality of the wastewater, the different local point sources, the physicochemical properties of particular organic compounds,


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and operational parameters of the wastewater treatment plant. Concentrations of organic pollutants are generally greater in industrial sewage than in domestic effluent (Bodzek and Janoszka, 1999; Smith, 2000). Over 300 organic chemicals from a diverse range of classes of compounds have been identified in biosolids and their concentrations vary from the pg kg 1 to the g kg 1 level ( Jacobs et al., 1987; Smith, 2000). Although water treatment plants were designed primarily to remove organic matter, and the mechanisms of degradation of bulk organic components are well studied and understood, the processes by which synthetic organics are degraded have received relatively little study. An organic contaminant could undergo a number of processes including (i) sorption to solid surfaces, (ii) volatilization, (iii) chemical degradation, and (iv) biodegradation. Generally, the more hydrophobic a compound is, the more susceptible to accumulation onto sewage sludge particles it will be. During primary sedimentation, hydrophobic contaminants may partition onto settled primary sludge solids. The tendency to accumulate in sludge solids can be assessed using the octonol–water partition coefficient (Kow) (Byrns, 2001). Contaminants with log Kow values less than 2.5 have low sorption potential and those with values greater than 4 have high sorption potential. Similarly, the tendency for volatilization can be gauged using Kow and Henry’s Law constant (Rogers, 1996). Although there is a scarcity of data on the behavior of organic contaminants during the water treatment and sludge digestion processes, some generalizations can be made (Rogers, 1996; Scow, 1982). For example, molecules with highly branched hydrocarbon chains are generally less susceptible to biodegradation than unbranched compounds and short chains are not as quickly degraded as long chains. In addition, unsaturated aliphatic compounds are generally more susceptible to degradation than saturated analogs. In general, due to their characteristically low water solubility and high lipophilicity, organic contaminants partition into sludges during sedimentation resulting in their accumulation in biosolids in concentrations several orders of magnitude greater than influent wastewater concentrations (Bhandari and Xia, 2005). Some organic contaminants are known, or are suspected to be endocrine disrupting chemicals (EDCs) and this has magnified interest in their presence in sewage, their fate during waste water treatment, and their possible presence in biosolids. The endocrine system is found in nearly all animals and is a complex network of glands that discharge hormones that regulate the body’s functions including growth, development, and maturation as well as the way various organs operate. EDCs possess the ability to alter or disrupt endocrine system function mimicking, antagonizing, or interfering with biosynthesis or biodegradation of endogenous hormones (Sonnenschein and Sato, 1998). Indeed, there is growing concern about the apparently increasing incidence of reproductive disorders and abnormal

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development in wildlife and reduced fertility in human males, problems that may be caused by EDCs that have been released anthropogenically into the environment (Ashby et al., 1997; Sonnenschein and Soto, 1998). A wide range of chemicals have been found, or are suspected to be capable of disrupting the endocrine system (Birkett, 2003a). These include: (i) persistent organochlorines and organohalogens (e.g., PCBs, dioxins, furans, brominated fire retardants), (ii) pesticides (e.g., DDT, atrazine, vinclozolin, TBT), (iii) alkyl phenols (e.g., nonylphenol, octylphenol), (iv) phytoestrogens (e.g., isoflavoids, lignins, B-sitosterol), and (v) natural and synthetic hormones (e.g., b-estradiole, ethynylestradiol). The last group of chemicals (above) is known as estrogenic EDCs and has received particular attention in recent times. Estrogens are a group of female sex hormones involved in the estrous cycle. They are excreted in urine and eliminated in feces and both naturally occurring estrogens and xenoestrogens (man-made analogs) have been identified in sewage, biosolids, and waste water treatment plant effluents (Birkett, 2003b). Analysis of organic contaminants in biosolids presents a number of challenges. The organic molecules are physically and/or chemically bound to the biosolids solid phase matrix and must be extracted with relatively harsh reagents/methods prior to analysis. Because extraction yields a complex mixture of organic compounds, the various organic fractions need to be segregated (cleaned) prior to analysis. Furthermore, concentrations of contaminants are often low, and sometimes below the limits of current analytical techniques. As a result, preconcentration is often necessary prior to analysis. Traditional methods entail soxhlet extraction, concentration using rotary evaporation and cleanup by column chromatography (Rogers, 1996). Contemporary methods, often preferred today, involve pressurized liquid extraction (PLE) followed by combined cleanup and concentration using solid phase extraction (SPE) (Cirelli et al., 2008; Jones-Lepp and Stevens, 2007). Quantification of organic compounds present is performed by one of a number of powerful chromatographic methods such as high performance liquid chromatography (HPLC), gas chromatography (GC), liquid chromatography (LC), and superficial fluid chromatography (SFC) and these are often coupled with molecule analysis by mass spectrometry (MS). In recent times, as methods of extraction, cleanup and analysis have become more rapid, an increasing number of surveys have reported the concentrations of various organic contaminants in wastewater and biosolids samples. However, as yet, there are only a few studies where the fate of such contaminants during wastewater treatment processes has been reported. Similarly, very little information is available on the behavior and degradation of biosolids-borne organic contaminants in soils following land application.


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Organic contaminants certainly have the potential to adversely impact on the soil/crop/animal system receiving land application of biosolids (Bhandari and Xia, 2005). Nevertheless, the relative risk of organic chemicals in biosolids is generally considered to be minimal due to the relatively low concentrations present and the many transformations (especially biodegradation) that can occur in soils (Epstein, 2003).The European Union has proposed limit values for several organic contaminants (or groups of contaminants) in biosolids and several European countries (e.g., Sweden, Germany, Denmark) have enforced a number of them. The European Union limits are for sum of halogenated organic compounds, linear alkylbenzene sulfonates, di-(ethylhexyl) phthalate, nonylphenol and nonylphenol ethoxylates, polynuclear aromatic hydrocarbons, polychlorinated biphenyls, and polychlorinated dibenzo-p-dioxins and -furans. Below, the most frequently detected organic contaminants in biosolids are considered. Their origin in sewage, potential toxicity, fate during wastewater treatment, and persistence or otherwise in the soil following land application are considered.

6.1. Organic compounds present 6.1.1. Phthalic acid esters (PAEs) These materials are manufactured in large quantities and are used predominantly as plasticizers to make plastics more flexible (Bhandari and Xia, 2005). The large majority of phthalates are used to plasticize polyvinyl chloride (PVC) to produce products ranging from kitchen and bathroom flooring to medical tubing, toys, footwear, electrical cables, packaging, and roofing. Di-(2-ethylhexyl) phthalate (DEHP) is the most widely used phthalate. Other common PAEs include Di-n-octyl phthalate (DnOP), butylbenzyl phthalate (BBP), Di-n-butyl phthalate (DnBP), diethyl phthalate (DEP), and dimethyl phthalate (DMP). Phthalates are used in non-PVC applications such as paints, rubber products, adhesives, cosmetics and toiletries (e.g., nail polish, perfumes), epoxy resins, adhesives, and printing inks. The widespread industrial and domestic use of products containing phthalates results in large amounts being washed down drains and into sewage systems (Marttinen et al., 2003; Palmquist and Hanaeus, 2005). Some PAEs are considered to be EDCs and/or carcinogens (Birkett, 2003a). Indeed, DEHP is considered to be an EDC (Hoyer, 2001) and has been shown to cause malformations of the reproductive system in male rats (Gray et al., 1999). BBP and DBP have the strongest estrogenic potencies (Harris et al., 1997). A wide range of bacteria can degrade PAEs under both aerobic and anaerobic conditions (Staples et al., 1997). The extent of biodegradation during anaerobic digestion is apparently related to the size of the alkyl side chain and compounds with the larger C-8 side chain are much more

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resistant to microbial attack ( Jianlong et al., 2000; Staples et al., 1997). As a result, the higher molecular mass PAEs such as DEHP and DnOP are considerably more persistent to anaerobic microbial degradation than lower molecular mass compounds (e.g., DPP and DEP). DEHP is generally considered persistent during sewage treatment (especially with anaerobic sludge digestion) (Bhandari and Xia, 2005; Scrimshaw and Lester, 2003). Total concentrations of PAEs in biosolids typically range from 12 to 200 mg kg 1 (Amir et al., 2005c; Cai et al., 2007b; Gibson et al., 2005; Harrison et al., 2006; Marttinen et al., 2003; Oliver et al., 2005). DEHP is consistently the most abundant in biosolids usually accounting for 25–95% of total PAEs present (Amir et al., 2005c; Cai et al., 2007b; Gibson et al., 2005; Oliver et al., 2005). Other PAEs such as DnOP, BBP, DnBP, DEP, and DMP are commonly present in biosolids in low concentrations ( 4) are retarded at the root surface due to sorption while polar chemicals with low values (e.g., Kow < 0.5) are less able to cross hydrophobic lipid membranes (Collins et al., 2006; Duarte-Davidson and Jones, 1996). Water and solutes transported in the xylem may be sorbed by stem components and/or accumulate in shoots and leaves. The organic chemicals accumulate in shoots as a result of equilibration of the aqueous phase in plant shoots with xylem constituents and sorption of the compound onto lipophilic shoot solids (McFarlane, 1994). Plant transpiration streamflow rate and lipophilic solids content of the plant will greatly influence accumulation of compounds in shoots (Collins et al., 2006). The potential for substantial sorption of the compounds to stem components during xylem transport increases with increasing lipophilicity of the chemical (Briggs et al., 1983; McGrady et al., 1987). Plant species can vary greatly in their ability to absorb and translocate organic compounds (Collins et al., 2006). Mattina et al. (2000, 2003, 2004) have shown that among a wide range of crops, plants within the Cucurbitaceae family (e.g., Cucurbita pepo L, zucchini) are especially adept at uptake of soil-bound DDE and chlordane. Similarly, Hulster et al. (1994) showed that courgette and pumpkin absorb and translocate relatively more PCCD/ Fs than most other crops. It has been suggested that these plants release root exudates which mobilize organic compounds from the soil thus increasing their availability (Hulster et al., 1994; Mattina et al., 2000). It is important to note here that soil characteristics are also important factors and that plant uptake of organic compounds is usually inversely related to soil organic matter content (Beck et al., 1996; Ryan et al., 1988). This is because hydrophobic organic compounds with a high Kow are strongly adsorbed to organic matter particles. As a result, plant uptake of organic compounds (e.g., PAHs) from solution culture is characteristically much higher than from soils at the same concentration of compound in soil solution (Gao and Ling, 2006). Biosolids particles are about 50% organic matter, so addition of biosolids to soils will tend to increase sorption (and compounds will already be sorbed to biosolids particles). The desorption process of these compounds from solid phase into soil solution will restrict their uptake into plant roots. The availability of organic compounds in soil solution is, therefore, usually the primary restriction to plant uptake of compounds including chlorobenzenes, organochlorines, and PAHs (Gao and Ling, 2006; Mattina et al., 2003; Wang and Jones, 1994a). Indeed, organic compounds are usually present in biosolids in low concentrations and they are so strongly sorbed by the soil/biosolids matrix that they exhibit very low bioavailability and accumulate in the edible portion of food crops at extremely low concentrations (O’Connor, 1996). Even where root


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accumulation occurs, translocation to shoots is usually negligible due to their low water solubility and high Kow values. Indeed, shoot concentration factors (ratio of concentration of organic contaminant in shoots to that in soil solution) are often an order of magnitude less than the equivalent values for root concentration (Gao and Ling, 2006). 6.2.2.2. Uptake from leaves In general, accumulation of organic contaminants into aboveground vegetation from biosolids-amended soils is thought to be dominated by vegetative uptake of contaminated vapor from the surrounding air (Beck et al., 1996; O’Connor, 1996). This has been demonstrated to be the major uptake pathway into plant foliage for a range of organic compounds including PAHs, PCBs, and PCDD/Fs (Bohme et al., 1999; Simonich and Hites, 1994; Welsch-Pausch et al., 1995). Compounds with high volatility and lipophilicity will have the highest bioconcentration into foliage. Compounds with a Henry’s law constant above 1 10 4 are generally considered susceptible to volatilization (Duarte-Davidson and Jones, 1996). Once volatilized from the soil into the atmosphere, chemicals may subsequently diffuse into plant leaves via the cuticle or stomata. Chemicals entering the leaves will diffuse into intercellular air spaces and partition to aqueous and lipophilic phases of adjacent plant tissues (Collins et al., 2006). Lipophilic leaf tissues include the waxy cuticle as well as membrane lipids, storage lipids, resins, and essential oils. Lipophilicity of the organic compound is therefore important and compounds with a log Kow of greater than 4 and high volatility have the greatest potential for foliar uptake (Duarte-Davidson and Jones, 1996). Partition of volatilized chemicals between plant foliage and air has also been related to the octanol/air partition coefficient (KOA) and a linear relationship between log KOA and shoot concentrations of PAHs and chlorobenzenes has been noted (Kipopoulou et al., 1999; Wang and Jones, 1994a). Differences in cuticular permeabilities and leaf/air bioconcentration for volatile organic compounds between plant species has been related principally to leaf lipid content (Collins et al., 2006). Organic chemicals may also come in contact with foliage following deposition in association with dust, aerosols, or atmospheric particulate matter. Some of this may accrue from resuspension of soil/biosolids particles. More particularly, chemicals may come in contact with foliage following direct application (e.g., surface application of biosolids to pastures). Once in contact with the leaves, particle-bound organic chemicals may diffuse through the cuticle and become sorbed to lipophilic material or permeate into the leaf interior (Collins et al., 2006). The permeability of the cuticle to organic chemicals in solution has been observed to be linearly related to Kow and inversely related to its molecular size (Riederer et al., 2002). Uptake of chemicals from particulate deposits will, however, be

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more complex since it will also depend on the capability of the compound to desorb from particle surfaces. 6.2.2.3. Other roles of plants The presence of plants can be an important factor in increasing the rate of degradation of organic contaminants in soils (Alkorta and Garbisu, 2001; Pilon-Smits, 2005; Salt et al., 1998) including those amended with biosolids (Laturnus et al., 2007). As a result, phytoremediation, which uses plants and associated rhizosphere microorganisms to remove or transform organic contaminants in soils, is an emerging technology (Alkorta and Garbisu, 2001; Susarla et al., 2002). Phytoremediation has been successfully employed to treat soils contaminated with a range of organics including chlorinated solvents, aromatic compounds, surfactants, and explosives (e.g., trinitrotoluene, TNT) (Susarla et al., 2002). Plants release a range of carbonaceous compounds into the rhizosphere as root exudates. As a result, microbial densities are 1–4 orders of magnitude higher than the surrounding bulk soil and the community also has a greater range of metabolic capabilities (Salt et al., 1998). Particular plant species could promote degradation of specific organic compounds. For example, it has been suggested that some species preferentially release phenols capable of supporting PCB-degrading bacteria (Fletcher and Hegde, 1995). Rhizosphere microorganisms may also accelerate degradation by volatilizing organics such as PAHs (Alkorta and Garbisu, 2001). Some bacteria can release biosurfactants that may make hydrophilic pollutants more water soluble (Volkering et al., 1998) and plant exudates can contain lipophilic compounds that increase pollutant water solubility and/or promote biosurfactant-producing microbial populations (Siciliano and Germida, 1998). In addition, both plant roots and microorganisms can release enzymes into soils which are involved in degradation of organic pollutants (e.g., laccases, dehalogenases, nitroreductases, nitrilases, and peroxidases) (Schnoor et al., 1995; Wolfe and Hoehamer, 2003). Enhanced biodegradation of organics induced by rhizosphere microflora has been reported for a number of compounds including pentachlorophenol (Ferro et al., 1994), surfactants (Knabel and Vestal, 1992), TCE (Shim et al., 2000; Walton and Anderson, 1990), PAHs ( Joner et al., 2002), and phenanthrene (Corgie et al., 2004; Fang et al., 2001). It has been suggested that in biosolids-amended soils the above mechanisms are of minor importance because biosolids already contain high concentrations of labile organic compounds and a large, active microbial community (Laturnus et al., 2007). These workers suggested that improved porosity and aeration induced by the presence of growing plants favors microbial activity and their degradative activities in biosolidsamended soils. As already discussed, uptake of organic contaminants by plants is usually low, so phytoextraction as a mechanism of remediation of organic contaminants is not usually a viable option. However, uptake can contribute to


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phytoremediation. Following uptake, organic compounds have multiple fates. They may be transformed to less toxic compounds and bound in plant tissues in nonavailable forms, they may undergo partial or complete degradation (phytodegradation) or they may be volatilized (phytovolatilization) (Alkorta and Garbisu, 2001; Susarla et al., 2002). Phytodegradation occurs when plant enzymes act on compounds and catalyze them partially or fully, thus reducing their concentrations in plant tissue. This can occur in both the root and shoot tissue. Those chemicals where evidence of significant degradation has occurred include trichloroethylene, benzene, pyrene, and TNT (Hannink et al., 2002; Susarla et al., 2002). Phytovolatilization occurs when volatile organic pollutants with a high water solubility and vapor pressure are absorbed, translocated to the leaves and lost to the atmosphere as a gas. Examples of organics that can be volatilized from plants include the solvents benzene and TCE and the fuel additive MTBE (Collins et al., 2006; Ma and Burken, 2003; Pilon-Smits, 2005). In general, concentrations of organic pollutants in biosolids-amended soils will be considerably lower than those present in situations where phytoremediation might be considered a strategy to clean up a contaminated soil. Nonetheless, since biosolids are generally land applied to agricultural land, the presence of growing plants does seem likely to enhance the degradation of many biosolids-borne organic contaminants that are introduced to soils. 6.2.3. Animals When biosolids is applied to agricultural land the potential transport of toxic organic chemicals to human food products is of concern. The pathways causing most apprehension are in animal production systems because persistent, lipophilic compounds bioconcentrate in body fat (i.e., meat products) and fat-containing products (e.g., milk). The pathways through which biosolids-derived compounds enter the grazing animal, and factors affecting this entry, were discussed previously (see, Section 5.5). Since most organic compounds are not particularly mobile and do not accumulate in large amounts in plant tops, the biosolids–soil–plant–animal pathway is usually of minor importance. The greatest intake from herbage will occur when biosolids are applied directly to established pasture, biosolids adhere to the herbage and animals have immediate access. Nevertheless, as with heavy metals, soil/biosolids ingestion during periods when forage is sparse is likely to be the major pathway of entry of biosolids-derived organic compounds into the grazing animals (Fries, 1996). As noted previously, this typically occurs most during winter when forage availability is least. Of the many organic contaminants present in biosolids, lipophilic, halogenated hydrocarbons (e.g., halogenated biphenyls, chlorinated pesticides and hydrocarbons, and PCDD/Fs) are of primary concern since they are resistant to degradation and bioconcentrate in fat of animals and animal

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products (Fries, 1996). Among these, compounds with low degrees of halogenation are metabolized and do not accumulate but higher degrees of halogenation block metabolism and bioconcentration occurs (McLachlan, 1993; Shiu and Mackay, 1986). Other compounds commonly found in biosolids such as phthalates, PAHs, acid phenols, nitrosamines, volatile aromatics, and aromatic surfactants are metabolized and do not generally accumulate in animal tissues (Fries, 1996). Bioconcentration factors (the ratio of concentration of compound in a tissue or product to the concentration in the diet) for halogenated organics (e.g., DDE, dieldrin, hexachlorobenzene, heptachlor, PCBs, chlorodibenzo-p-dioxin, and chlorodibenzofuran) are usually in the range of 5–6 in milk fat of cows and the body fat of sheep and cattle (Fries and Marrow, 1975; Fries et al., 1969; Harrison et al., 1970; Parker et al., 1980; van den Hoek et al., 1975; Willett et al., 1990). To minimize bioaccumulation of halogenated hydrocarbons in grazing animals it is important to allow time for biosolids to be washed off the surface of pasture leaves prior to grazing and minimize the chances of soil ingestion during the winter months.

7. Synthesis and Conclusions Wastewater derived from domestic sources (human feces, urine, and graywater) plus that from commercial enterprises and industry, and that from runoff into stormwater is treated in municipal wastewater treatment plants in a series of processes primarily aimed at removing the dissolved and suspended organic material. Biosolids are by-products of wastewater treatment and consist of approximately 50% inorganic and 50% organic material. The organic components originate principally from two sources: (i) human feces settled out during primary treatment (sedimentation) and (ii) bacterial cells settled out during secondary treatment in which bacterial activity is used to remove the suspended/dissolved organic load from the primary treatment effluent. The organic components undergo a degree of degradation and humification particularly during sludge stabilization (often anaerobic digestion). The inorganic component mainly settles out during sedimentation and originates from sources such as local soil and sediments, broken glass, and inorganic residuals of food/feces (e.g., silica). Biosolids contain both inorganic and organic contaminants which originate from the influent sewage wastewater. The major inorganic contaminants are heavy metals (e.g., Cu, Zn, Cd, Pb, Ni, Cr, As) which mainly originate from discharges from industry and from domestic graywater including leaching from Cu and Pb pipes and Zn from domestic products (skin creams, deodorants, etc.). Heavy metals enter the sludge during primary treatment through their association/adsorption to sedimenting


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particles and during secondary treatment through adsorption to bacterial cell walls and/or accumulation into bacterial cells. Concerns regarding the heavy metal loads in biosolids have resulted in guidelines and regulations being developed in many parts of the world which are usually based on maximum allowable metal concentration limits in biosolids and/or the allowable loading limits of metals added in biosolids to the soil. A limitation of these approaches is that they consider total rather than biologically-active (extractable) concentrations of heavy metals in soils. Where agronomic rates are used (2–8 Mg ha 1), heavy metal toxicities limiting crop growth in biosolids-amended soils are very rare. Nevertheless, a major concern is that heavy metals may accumulate in edible portions of plants and subsequently enter the food chain and have toxic effects on grazing animals and/or humans ingesting them. For the most part, heavy metals are not readily translocated to the aboveground edible portions of crops so toxicities from ingestion of food crops are not likely under current regulations. The main potential pathway for accumulation of heavy metals into the meat of grazing animals is via direct soil/biosolids ingestion and substantial accumulation is unlikely under adequate grazing management. There is a growing body of evidence that biosolids-induced heavy metal accumulation in soils can have negative effects on soil microbial/biochemical activity but the significance/ importance of this has yet to be fully understood. With recent improvements in extraction techniques and analytical chromatographic methods, there is an increasing body of research involving surveys of organic contaminants in biosolids. These include PAEs, PAHs, PCBs, chlorobenzenes, chlorophenols, dioxins, organotins, pesticides, brominated flame retardants, surfactants, pharmaceuticals, and natural and synthetic hormones. Due to their low water solubility and high lipophilicity they are generally believed to partition into sludges during sedimentation. Nonetheless, the fate of these chemicals during wastewater treatment processes is not well characterized. Many are not readily degradable under anaerobic conditions and therefore persist in biosolids following anaerobic digestion. There is very limited information in the literature on the behavior and fate of biosolids-borne organic contaminants in agricultural soils. However, the literature available suggests some are rapidly lost following land application since they are readily degraded under aerobic soil conditions, lost by volatilization or by photolysis. Some are, however, known or thought to be only slowly degraded (e.g., organotins, brominated flame retardants, and high molecular mass PAHs) and some are highly persistent (e.g., dioxins and organochlorines). Uptake of organic contaminants from soil into plants is characteristically low and the main pathway is via volatilization from the soil surface and then uptake of contaminated air by aboveground vegetation. Accumulation of organics into grazing animals could occur through ingestion of biosolids adhering to vegetation directly after is application and by direct soil/biosolids ingestion. Many organics are

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metabolized in animals but halogenated biphenyls, chlorinated pesticides and hydrocarbons, and PCDD/Fs are of primary concern since they are resistant to degradation and bioconcentrate in animal fat and animal products (e.g., milk). The European Union has proposed limit values for several organic contaminants/groups of contaminants in biosolids although the relevance of these is not yet fully understood. A better understanding of the reactions and fate of organic contaminants in soils is required before more rigorous regulations can be drafted. If sewage entering a wastewater treatment plant contains a substantial loading of heavy metals or organic contaminants, it is very likely that biosolids produced at the plant will also have a substantial contaminant load. Thus, to lower contaminant levels in biosolids, reductions in the release of contaminants into the sewer system are required. The most practicable way of doing this is at point source outlets to the sewer. Indeed, in recent years, enforcement of tighter regulations has greatly reduced industrial inputs of heavy metals to wastewater streams since industry is expected to pretreat their wastewater (where appropriate) prior to its release. In addition, the use of extremely persistent organics such as dioxins and organochlorines has been effectively terminated making contamination of wastewater with such compounds less pronounced. As a result, contaminants in biosolids are becoming less of an issue and the positive effects of recycling the biosolids via land application onto agricultural or forestry lands are gaining momentum. Indeed, land application is generally seen to be the most economical and beneficial way to deal with biosolids. They contain high concentrations of N, P, Ca, and Mg but K is usually low and needs to be supplemented for crop production. Biosolids applications need to be based on the N needs of crops to be grown so that excessive applications can be avoided thus minimizing leaching and gaseous losses of N to the surrounding environment. Because the bulk of N in biosolids is in organic form, an estimate of N mineralization potential is required to calculate effective N rates. The P content of biosolids is high relative to the plant requirement (based on optimum N supply) and care needs to be taken to avoid runoff losses of P and/or P leaching if available P becomes very high. Where available soil micronutrient levels are low, supply of micronutrients such as Cu, Zn, Fe, and Mn in biosolids may be important. Biosolids applications have other positive effects. Addition of partly humified organic matter can increase soil organic matter content, thereby improving soil physical conditions (e.g., increased porosity and aggregation) and increasing the size and activity of the soil microbial biomass and soil enzyme activity. Areas where a better understanding is required and where research needs to be concentrated include:


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(1) Development of appropriate measures of the biologically active pool of heavy metals in biosolids-amended soils and development of guidelines using such values. Continued monitoring of long-term field sites using such measures is needed to understand long-term effects. (2) A better understanding of the negative effects of sludge-borne heavy metals on soil microbial and biochemical activity and of the significance of this to the soil system. (3) A better understanding of the fate of organic contaminants in sewage during wastewater treatment processes. (4) An in-depth understanding of the fate of organic contaminants in soils (and the mechanisms of their degradation) following land application of contaminated biosolids. An understanding of the effects of accumulated organic contaminants on soil processes is also needed. (5) Development of scientifically based critical concentrations/loadings for organic contaminants.

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Index

A Agricultural production systems SIMultor (APSIM) legume–cereal rotations systems, 115 SOILN module, 114 water-use efficiency (WUE) and N-use efficiency (NUE) assesment, 115–116 wheat–chickpea rotation trial, 116 Arsenic (As) aqueous chemistry, 139 bioavailability assessment chemical forms, 37–39 factors, 36–37 modification, 39 reliable relative bioavailability (RBA), 40–41 risk assessment, 39–40 soil, 35–36 biogeochemical processes arsenic desorption, anaerobiosis, 150–154 microbial arsenate reduction, 149–150 concentrations controlling processes, South (east) Asia biogeochemical and hydrologic process, 156 microbial reduction of, 155 oxidative weathering, 154 reductive release of, 155–156 desorption, soils and sediments comparative desorption, 147–149 displacement and mobilization, 146–147 oxidative dissolution, 146 pollution and hazardous concentrations, 138–139 surface and solid phases adsorption envelopes, As(V) and As(III), 143–144 arsenate precipitation, 145 arsenite vs. arsenate, 142–145 humic substances adsorption, 142 retention of, 139–142 sulfidogenesis, 145–146 B Barley-based rotation trials, 104 Biodegradation branched hydrocarbon chains, 209 brominated flame retardants, 222

chlorobenzenes (CBs), 215 chlorophenols, 218 dioxin, 220 estrogenic organics, 225 linear alkylbenzene sulfonate (LAS), 224 phthalic acid esters (PAEs), 211–212 polybrominated diphenyl ethers (PBDEs), 223 polychlorinated biphenyls (PCBs), 216 Biogeochemical processes arsenic desorption, anaerobiosis, 150–154 microbial arsenate reduction, 149–150 Biological metal uptake children ingestion of, 13–14 food-chain transfer and risks, 14–16 soil ingestion risks, 12–13 Biosolids heavy metal contaminants animal ingestion, 207–208 definition, 182 extractable fractions, 185–187 plant response and metal uptake, 202–207 soil application, 187–202 total concentrations, 183–185 inorganic components ash content, 174–175 X-ray fluorescence analysis, 175 nutrient content and release calcium (Ca) and magnesium (Mg), 181 electrical conductivity (EC), soil, 182 micronutrients, 181–182 nitrogen (N), 175–179 pH change, soil, 182 phosphorus (P), 179–181 potassium (K), 181 organic contaminants branched hydrocarbon chains, 209 brominated flame retardants, 222–223 chlorobenzenes (CBs), 214–215 chlorophenols, 217–218 endocrine disrupting chemicals (EDCs), 209–210 European union limits, 211 natural and synthetic hormones, 225–226 octonol–water partition coefficient (Kow), 209 organochlorine pesticides (OCPs), 216–217 organotin compounds, 220–222 pharmaceuticals and personal care products, 225

269

270

Index

Biosolids (cont.) phthalic acid esters (PAEs), 211–212 polychlorinated biphenyls (PCBs), 215–216 polychlorinated dibenzodioxins (PCDDs)/ dibenzofurans (PCDFs), 218–220 polycyclic aromatic hydrocarbons (PAHs), 213–214 potential transfer to groundwater, plants, and animals, 227–234 presence and level, 208–209 surfactants, 223–224 organic matter characterization, humic substances, 170–171 composition, 169–170 composting, 171 humification, 170 soil application, 171–174 water-soluble labile portion, 171 sewage treatment processes dewatering method, 169 primary and secondary treatment, 168 stabilization methods, 169 thickening, sludge, 168–169 C CERES-wheat crop simulation model, 116–117 Chlorobenzenes (CBs), 214–215 Chlorophenols, 217–218 Conservation/compost tillage trial, 104–105 E Endocrine disrupting chemicals (EDCs), 209–210

extractable fractions available and unavailable forms, 186–187 chemical speciation, 185–186 cocomposting method, 187 sequential fractionation schemes, 186 plant response and metal uptake accumulation of, 205–207 metal toxicity and tolerance, 202–205 phytotoxicity, 202 soil application biosolids property effect, 189–191 extraction methods of, 187–189 metal availability over time, 193–194 microbial/biochemical effects, 198–202 mobility and leaching of, 194–198 soil property effect, 191–193 total concentrations effluent source, 183 guidelines and regulations, 185 waste water treatment plants, 183–185 I Integrated cropping systems barley-based rotation trials, 104 conservation/compost tillage trial, 104–105 cropping systems productivity trial grain and straw yields and quality, 97–99 soil nitrogen dynamics, 99–100 soil organic matter, 100–101 water-use efficiency, 99 crop sequence effect, 95 grazing management rotation trial, 101–104

F

L

Fertilizers, nitrogen rainfed crops response genetic differences, crops, 79–80 Morocco and the Maghreb countries, 73–79 soil organic matter, 68–69 West Asian countries, 70–73 usage trend, 66–67 Food-chain transfer metals, 14–16 plants, organic chemicals leaves uptake, 231–232 rhizosphere activity, 232–233 root uptake and translocation, 228–231

Leaching, heavy metals, 194–198 Lead (Pb), bioavailability assessment absorption, 29 bioaccessibility, soil Pb, 34–35 blood level reduction, children, 29–30 soil/dust Pb chemical speciation, 30 chloropyromorphite (CP) formation, 33–34 feeding tests, 32–33 RBA assessment, 30–32 source, 28

G Grazing management rotation trial, 101–104 H Heavy metal contaminants, biosolids animal ingestion, 207–208 definition, 182

M Mediterranean agroecosystems climate rainfall sketch and intensity, 59–60 seasonal rains and temperature effect, 58 crops and farming systems agroecological condition, 62 barley production, 61–62 niche crops production, 62 wheat production, 61

271

Index

fertilizer use trends, 66–67 integrated cropping systems barley-based rotation trials, 104 conservation/compost tillage trial, 104–105 cropping systems productivity trial, 96–101 crop sequence effect, 95 grazing management rotation trial, 101–104 land features, 60–61 nitrogen, 65–66 nitrogen fixation, dryland conditions food legumes, 86 pasture and forage legumes, 89–91 rhizobia, inoculation, and cultivar interactions, 86–89 rainfed crops response to genetic differences, crops, 79–80 Morocco and the Maghreb countries, 73–79 soil organic matter, 68–69 West Asian countries, 70–73 soil status assessment Cate–Nelson graphical method, 81 Kjeldahl method, 81 mineralization indices, 83–84 mineral N tests, 82 nitrate test, 81–82 plant analysis, 84–85 Metal(Loid)s, bioavailability assessment arsenic chemical forms, 37–39 factors affect, 36–37 modification, 39 reliable relative bioavailability (RBA), 40–41 risk assessment, 39–40 soil, 35–36 biological metal uptake children ingestion of, 13–14 food-chain transfer and risks, 14–16 soil ingestion risks, 12–13 chemistry equilibrium, soil, 22–25 speciation in, 25–28 extractability, availability prediction chelation methods, 21–22 diffusive gradients in thinfilms method, 22 in vitro bioaccessibility, 16–21 lead (Pb) absorption, 29 bioaccessibility, soil Pb, 34–35 blood level reduction, children, 29–30 soil/dust Pb, 30–34 source, 28 soil risks bioavailability and soil element, 7–8 microbes and fauna toxicity, 11–12 phytotoxicity from, 8–11 Micellization, 224

N Neutron activation analysis (NAA), metal bioaccessibility, 17–18 Nitrogen (N) biosolids agronomic biosolids rate, 178–179 agronomic response, 177 ammonia volatilization, 177 mineralization, 176–177 fertilizer and rainfed crops response field responses, 69–79 genetic differences, crops, 79–80 soil organic matter, 68–69 fixation, mediterranean dryland conditions food legumes, 86 pasture and forage legumes, 89–91 rhizobia, inoculation, and cultivar interactions, 86–89 integrated cropping systems barley-based rotation trials, 104 conservation/compost tillage trial, 104–105 cropping systems productivity trial, 96–101 crop sequence effect, 95 grazing management rotation trial, 101–104 mediterranean region, 65–66 modelling systems APSIM usage, 115–116 arid environment models, 118–120 CERES-wheat model, 117 crop–soil simulation models, 114 CropSyst model, 118 root zone water quality model (RZWQM), 117 SOILN module, 114–115 potential losses, dryland cropping ammonia volatilization, 93–94 environmental concerns, 92 loss mechanisms, 92–93 urea fertilizers effect, 94–95 rainfed environments, 64–65 soil status assessment Cate–Nelson graphical method, 81 Kjeldahl method, 81 mineralization indices, 83–84 mineral N tests, 82 nitrate test, 81–82 plant analysis, 84–85 supplemental irrigation systems seasonal weather effect, 106 wastewater usage, 106–107 yeild responses, 105–106 tracer (15N), rainfed cropping systems N fixation estimation, grain legumes, 111–113 rotation effect implications, 110–111 wheat, cereal–legume rotation, 108–110

272

Index O

Organic contaminants, biosolids branched hydrocarbon chains, 209 brominated flame retardants applications, 222 biodegradation, 223 source, 222–223 chlorobenzenes (CBs) concentrations in, 214–215 derivatives and applications, 214 land application, 215 chlorophenols atmospheric deposition, 218 pentachlorophenol (PCP), 217–218 endocrine disrupting chemicals (EDCs), 209–210 European union limits, 211 natural and synthetic hormones, 225–226 octonol–water partition coefficient (Kow), 209 organochlorine pesticides (OCPs), 216–217 organotin compounds biodegradation, 221 classes and apllication, 220 concentrations in, 221 source, 220–221 pharmaceuticals and personal care products, 225 phthalic acid esters (PAEs) applications, 211 microbial degradation, 211–212 polychlorinated biphenyls (PCBs) applications, 215 atmospheric deposition and soil application, 216 polychlorinated dibenzodioxins (PCDDs)/ dibenzofurans (PCDFs) biodegradation, 220 source, 219–220 toxicity, 219 polycyclic aromatic hydrocarbons (PAHs) land application, 214 source and lipophilicity, 213 potential transfer to groundwater, plants, and animals, 227–234 presence and level, 208–209

surfactants linear alkylbenzene sulfonate (LAS), 223–224 micellization, 224 properties, 223 Organochlorine pesticides (OCPs), 216–217 Organotin compounds, 220–222 P Phosphorus (P), biosolids phytoavailability, 180–181 soil accumulation, 180 surface runoff, 181 Phthalic acid esters (PAEs), 211–212 Phytotoxicity, metal effect Ni–Al layered double hydroxides (LDH), 9 Ni-induced Fe deficiency symptoms, 10 precipitation and synchrotron XRF (S-XRF) analysis, 11 Polychlorinated biphenyls (PCBs), 215–216 Polychlorinated dibenzodioxins (PCDDs)/ dibenzofurans (PCDFs), 218–220 Polycyclic aromatic hydrocarbons (PAHs), 213–214 R Reliable relative bioavailability (RBA) arsenic (As) (see Arsenic (As), bioavailability assessment) lead (Pb) (see Lead (Pb), bioavailability assessment) Root zone water quality model (RZWQM), 117 S Sewage treatment processes biosolids (see Biosolids) dewatering method, 169 primary and secondary treatment, 168 stabilization methods, 169 thickening, sludge, 168–169 Soil-N routine/module (SOILN), 114–115 Sulfidogenesis, 145–146 Surfactants, 223–224

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