Reviews of Environmental Contamination and Toxicology Volume 178

Reviews of Environmental Contamination and Toxicology VOLUME 178

Reviews of Environmental Contamination and Toxicology Continuation of Residue Reviews

Editor

George W. Ware Editorial Board Lilia A. Albert, Xalapa, Veracruz, Mexico D.G. Crosby, Davis, California, USA ⴢ Pim de Voogt, Amsterdam, The Netherlands O. Hutzinger, Bayreuth, Germany ⴢ James B. Knaak, Getzville, NY, USA Foster L. Mayer, Gulf Breeze, Florida, USA ⴢ D.P. Morgan, Cedar Rapids, Iowa, USA Douglas L. Park, Washington DC, USA ⴢ Ronald S. Tjeerdema, Davis, California, USA Raymond S.H. Yang, Fort Collins, Colorado, USA Founding Editor Francis A. Gunther

VOLUME 178

Coordinating Board of Editors DR. GEORGE W. WARE, Editor Reviews of Environmental Contamination and Toxicology 5794 E. Camino del Celador Tucson, Arizona 85750, USA (520) 299-3735 (phone and FAX) DR. HERBERT N. NIGG, Editor Bulletin of Environmental Contamination and Toxicology University of Florida 700 Experimental Station Road Lake Alfred, Florida 33850, USA (941) 956-1151; FAX (941) 956-4631 DR. DANIEL R. DOERGE, Editor Archives of Environmental Contamination and Toxicology 6022 Southwind Drive N. Little Rock, Arkansas, 72118, USA (870) 791-3555; FAX (870) 791-2499

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Foreword

International concern in scientific, industrial, and governmental communities over traces of xenobiotics in foods and in both abiotic and biotic environments has justified the present triumvirate of specialized publications in this field: comprehensive reviews, rapidly published research papers and progress reports, and archival documentations. These three international publications are integrated and scheduled to provide the coherency essential for nonduplicative and current progress in a field as dynamic and complex as environmental contamination and toxicology. This series is reserved exclusively for the diversified literature on “toxic” chemicals in our food, our feeds, our homes, recreational and working surroundings, our domestic animals, our wildlife and ourselves. Tremendous efforts worldwide have been mobilized to evaluate the nature, presence, magnitude, fate, and toxicology of the chemicals loosed upon the earth. Among the sequelae of this broad new emphasis is an undeniable need for an articulated set of authoritative publications, where one can find the latest important world literature produced by these emerging areas of science together with documentation of pertinent ancillary legislation. Research directors and legislative or administrative advisers do not have the time to scan the escalating number of technical publications that may contain articles important to current responsibility. Rather, these individuals need the background provided by detailed reviews and the assurance that the latest information is made available to them, all with minimal literature searching. Similarly, the scientist assigned or attracted to a new problem is required to glean all literature pertinent to the task, to publish new developments or important new experimental details quickly, to inform others of findings that might alter their own efforts, and eventually to publish all his/her supporting data and conclusions for archival purposes. In the fields of environmental contamination and toxicology, the sum of these concerns and responsibilities is decisively addressed by the uniform, encompassing, and timely publication format of the Springer-Verlag (Heidelberg and New York) triumvirate: Reviews of Environmental Contamination and Toxicology [Vol. 1 through 97 (1962–1986) as Residue Reviews] for detailed review articles concerned with any aspects of chemical contaminants, including pesticides, in the total environment with toxicological considerations and consequences. Bulletin of Environmental Contamination and Toxicology (Vol. 1 in 1966) for rapid publication of short reports of significant advances and discoveries in the fields of air, soil, water, and food contamination and pollution as well as v

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methodology and other disciplines concerned with the introduction, presence, and effects of toxicants in the total environment. Archives of Environmental Contamination and Toxicology (Vol.1 in 1973) for important complete articles emphasizing and describing original experimental or theoretical research work pertaining to the scientific aspects of chemical contaminants in the environment. Manuscripts for Reviews and the Archives are in identical formats and are peer reviewed by scientists in the field for adequacy and value; manuscripts for the Bulletin are also reviewed, but are published by photo-offset from cameraready copy to provide the latest results with minimum delay. The individual editors of these three publications comprise the joint Coordinating Board of Editors with referral within the Board of manuscripts submitted to one publication but deemed by major emphasis or length more suitable for one of the others. Coordinating Board of Editors

Preface

Thanks to our news media, today’s lay person may be familiar with such environmental topics as ozone depletion, global warming, greenhouse effect, nuclear and toxic waste disposal, massive marine oil spills, acid rain resulting from atmospheric SO2 and NOx, contamination of the marine commons, deforestation, radioactive leaks from nuclear power generators, free chlorine and CFC (chlorofluorocarbon) effects on the ozone layer, mad cow disease, pesticide residues in foods, green chemistry or green technology, volatile organic compounds (VOCs), hormone- or endocrine-disrupting chemicals, declining sperm counts, and immune system suppression by pesticides, just to cite a few. Some of the more current, and perhaps less familiar, additions include xenobiotic transport, solute transport, Tiers 1 and 2, USEPA to cabinet status, and zerodischarge. These are only the most prevalent topics of national interest. In more localized settings, residents are faced with leaking underground fuel tanks, movement of nitrates and industrial solvents into groundwater, air pollution and “stay-indoors” alerts in our major cities, radon seepage into homes, poor indoor air quality, chemical spills from overturned railroad tank cars, suspected health effects from living near high-voltage transmission lines, and food contamination by “flesh-eating” bacteria and other fungal or bacterial toxins. It should then come as no surprise that the ‘90s generation is the first of mankind to have become afflicted with chemophobia, the pervasive and acute fear of chemicals. There is abundant evidence, however, that virtually all organic chemicals are degraded or dissipated in our not-so-fragile environment, despite efforts by environmental ethicists and the media to persuade us otherwise. However, for most scientists involved in environmental contaminant reduction, there is indeed room for improvement in all spheres. Environmentalism is the newest global political force, resulting in the emergence of multi-national consortia to control pollution and the evolution of the environmental ethic. Will the new politics of the 21st century be a consortium of technologists and environmentalists or a progressive confrontation? These matters are of genuine concern to governmental agencies and legislative bodies around the world, for many serious chemical incidents have resulted from accidents and improper use. For those who make the decisions about how our planet is managed, there is an ongoing need for continual surveillance and intelligent controls to avoid endangering the environment, the public health, and wildlife. Ensuring safety-

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in-use of the many chemicals involved in our highly industrialized culture is a dynamic challenge, for the old, established materials are continually being displaced by newly developed molecules more acceptable to federal and state regulatory agencies, public health officials, and environmentalists. Adequate safety-in-use evaluations of all chemicals persistent in our air, foodstuffs, and drinking water are not simple matters, and they incorporate the judgments of many individuals highly trained in a variety of complex biological, chemical, food technological, medical, pharmacological, and toxicological disciplines. Reviews of Environmental Contamination and Toxicology continues to serve as an integrating factor both in focusing attention on those matters requiring further study and in collating for variously trained readers current knowledge in specific important areas involved with chemical contaminants in the total environment. Previous volumes of Reviews illustrate these objectives. Because manuscripts are published in the order in which they are received in final form, it may seem that some important aspects of analytical chemistry, bioaccumulation, biochemistry, human and animal medicine, legislation, pharmacology, physiology, regulation, and toxicology have been neglected at times. However, these apparent omissions are recognized, and pertinent manuscripts are in preparation. The field is so very large and the interests in it are so varied that the Editor and the Editorial Board earnestly solicit authors and suggestions of underrepresented topics to make this international book series yet more useful and worthwhile. Reviews of Environmental Contamination and Toxicology attempts to provide concise, critical reviews of timely advances, philosophy, and significant areas of accomplished or needed endeavor in the total field of xenobiotics in any segment of the environment, as well as toxicological implications. These reviews can be either general or specific, but properly they may lie in the domains of analytical chemistry and its methodology, biochemistry, human and animal medicine, legislation, pharmacology, physiology, regulation, and toxicology. Certain affairs in food technology concerned specifically with pesticide and other food-additive problems are also appropriate subjects. Justification for the preparation of any review for this book series is that it deals with some aspect of the many real problems arising from the presence of any foreign chemical in our surroundings. Thus, manuscripts may encompass case studies from any country. Added plant or animal pest-control chemicals or their metabolites that may persist into food and animal feeds are within this scope. Food additives (substances deliberately added to foods for flavor, odor, appearance, and preservation, as well as those inadvertently added during manufacture, packing, distribution, and storage) are also considered suitable review material. Additionally, chemical contamination in any manner of air, water, soil, or plant or animal life is within these objectives and their purview.

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Normally, manuscripts are contributed by invitation, but suggested topics are welcome. Preliminary communication with the Editor is recommended before volunteered review manuscripts are submitted. Tucson, Arizona

G.W.W.

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Table of Contents

Foreword ....................................................................................................... Preface ..........................................................................................................

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Influence of Aging on Metal Availability in Soils ..................................... KOEN LOCK AND COLIN R. JANSSEN

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Algal Toxicity Tests for Environmental Risk Assessment of Metals ........ COLIN R. JANSSEN AND DAGOBERT G. HEIJERICK

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Chemistry of Chromium in Soils with Emphasis on Tannery Waste Sites ................................................................................................... S. AVUDAINAYAGAM, M. MEGHARAJ, G. OWENS, R.S. KOOKANA, D. CHITTLEBOROUGH, AND R. NAIDU Chromium—Microorganism Interactions in Soils: Remediation Implications ............................................................................ SARA PARWIN BANU KAMALUDEEN, MALLAVARAPU MEGHARAJ, ALBERT L. JUHASZ, NABRATTIL SETHUNATHAN, AND RAVI NAIDU

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Index ............................................................................................................. 165

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 Springer-Verlag 2003

Rev Environ Contam Toxicol 178:1–21

Influence of Aging on Metal Availability in Soils Koen Lock and Colin R. Janssen Contents I. Introduction ............................................................................................................ II. Environmental Availability ................................................................................... A. Soil Components ............................................................................................... B. Natural Soils ...................................................................................................... III. Environmental Bioavailability ............................................................................... A. Plants ................................................................................................................. B. Invertebrates ...................................................................................................... IV. Toxicological Bioavailability ................................................................................ A. Plants ................................................................................................................. B. Invertebrates ...................................................................................................... C. Microorganisms ................................................................................................. V. Conclusions and Recommendations ...................................................................... Summary ...................................................................................................................... References ....................................................................................................................

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I. Introduction Considerable evidence is available showing that the total metal concentration in a soil is not a good predictor of its availability (Van Gestel 1997; Van Straalen and Denneman 1989). Consequently, risk assessments and soil quality objectives based on this type of information may not be protective for the terrestrial ecosystem. Bioavailability is a critical concept in ecotoxicology because substances that are not bioavailable, by definition, are unlikely to cause adverse effects because organisms are not exposed. As this term is used in different contexts, Landrum et al. (1994) proposed the following definitions: • Environmental availability is the proportion of the chemical in the environment (or a specified part) that can be involved in a particular process and is subject to all physical, chemical, and biological modifying influences. • Environmental bioavailability is the fraction of the environmentally available material that an organism accumulates when processing or encountering a given environmental medium. • Toxicological bioavailability is the fraction of the exposure dose or concentra-

Communicated by George W. Ware. K. Lock ( ), C.J. Janssen Ghent University, Laboratory of Environmental Toxicology and Aquatic Ecology, J. Plateaustraat 22, B-9000 Ghent, Belgium.

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tion that is absorbed and/or adsorbed by an organism, distributed by the systemic circulation, and ultimately presented to the receptors or sites of toxic action. As a result of long-term chemical processes, the bioavailability of metals in soils can decrease with time, with little or no reduction in the total metal concentration as determined by procedures that rely on vigorous extractions. Hence, total metal concentration measurements are often not relevant for prediction of potential exposure to, and thus risk from, contaminated soils. Although the effects of these long-term processes, i.e., aging, have been frequently reported, most regulatory frameworks are currently not making use of the available information. Therefore, a more widespread recognition of the influence of aging on metal bioavailability in soils is necessary among scientists, environmental engineers, regulators, and the public at large. Safe environmental concentrations for metals in the terrestrial compartment, such as Predicted No Effect Concentrations and Maximum Permissible Concentrations, are usually derived using the results of ecotoxicity tests in which freshly spiked soils are used. However, environmental quality standards calculated on the basis of this kind of data can overlap considerably with natural background concentrations (Lock and Janssen 2001a). This practice indicates that exposure, and thus risk, is probably overestimated when hazardous concentrations are based on this type of data. Incorporation of the effect of aging in the environmental risk assessment of metal-contaminated soils may contribute to a more realistic assessment of the impact of metals on terrestrial ecosystems. However, this is only possible when enough data are available to quantify the aging effect. In this review, the current knowledge about the influence of aging on metal bioavailability is reviewed. The chapter focuses on the effect of aging on metal bioavailability and the consequences for environmental risk assessment. No attempt was made to summarize data about the mechanisms of the aging process on a chemical basis. The paper is structured according to the distinction between the three different definitions of availability as proposed by Landrum et al. (1994).

II. Environmental Availability A. Soil Components As natural soils are very complex, a lot of experimental work on metal aging has been conducted using only soil components such as metal oxyhydroxides, calcites, and clays. Usually, batch experiments are conducted to study changes in metal availability with time. For Cu, Pb, Ni, Zn, and Cr(III), but not for Cd, a measurable fraction of the bound metal was not easily desorbed from ferrihydrite. This fraction increased with increasing pH and duration of the high-pH stage and increased more or less continuously in sequential cycles. With increasing total metal concentrations in the system, the fraction of the adsorbed metal that slowly desorbed decreased.

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Slowly reversible adsorption increased the longer the system was held at alkaline pH (Schultz et al. 1987). A gradual interchange of some readily desorbed copper into a fraction that was not readily desorbed from goethite was observed after an initial time lag of 4–6 d between adsorption and desorption (Padmanabham 1983). Long-term changes (400 d) in Cu activity were also observed in an iron oxide system (McBride et al. 1998), and total dissolved Cu decreased in a ferric oxide system after 200 d of aging (Martinez and McBride 1999). Decreased Zn solubility was indicated after 200 d of aging, and a further decrease occurred after the suspensions with ferric oxide were heated for 60 d at 70°C. However, Cd solubility remained approximately constant during the 200 d of aging at room temperature (Martinez and McBride 1998). The adsorption of Ni, Zn, and Cd to goethite increased with pH, reaction time, and temperature. Relative adsorption decreased with increasing initial metal concentration. After 42 d of reaction time, none of the metals had completely reached adsorption equilibrium (Barrow et al. 1989; Bruemmer et al. 1988). The pH sorption curves were displaced to lower pH values as the reaction time increased from 2 to 1008 hr. Similar effects were caused by a change in temperature from 5° to 35°C. The sorption of Ni, Cd, and Zn by goethite occurred by an initial fast step and a subsequent slow reaction that continued for several weeks (Gerth et al. 1993). With increasing residence times up to 1 yr, precipitated phases of gibbsite increased in stability, as shown by decreasing amounts of Ni release as effected by HNO3 and ethylenediaminetetraacetic acid (EDTA) treatments (Scheckel and Sparks 2000; Scheckel et al. 2000). Hydrous ferric oxide aged without the metal ions displayed hysteresis between the adsorption–desorption curves, and no substantial shifts in fractional metal adsorption were observed with pH throughout 21 wk of aging. Hydrous ferric oxide aged with metal ions displayed increasing desorption hysteresis with time for Co and Cd but not for Pb (Ainsworth et al. 1994). In neither the presence nor the absence of added lead ions was the aging effect in ferric oxide very pronounced even after several weeks (Gadde and Laitinen 1974). Mn oxides sorbed larger amounts of both Co and Cd than did the Fe oxides. Increasing the sorption period from 1 wk to 12–15 wk did not noticeably increase the amounts of Co or Cd sorbed by the oxides. However, increasing the sorption period had a marked effect on the subsequent rate of desorption of the sorbed metals (Backes et al. 1995). With time and temperature, the amount of Cd and Zn sorbed to the hydrous oxides of Al, Fe, and Mn gradually increased (Trivedi and Axe 2000). Concurrent sorption of phosphate, Zn, and Cd by a hydrous Fe oxide sample was investigated by enclosing sparingly soluble Zn or Cd phosphate in dialysis tubings and equilibration with the oxide for up to 2 yr. The hydrous Fe oxide was capable of retaining Zn and Cd, but there was no complementary effect of metal sorption on phosphate sorption and vice versa (Kuo 1986). The aforementioned data clearly indicate that, in batch experiments with metal oxyhydroxides, metal availability decreases with time. However, for metals with a great ionic radius, aging seems to have less effect on metal availability. Furthermore, several environmental parameters such as pH,

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temperature, and total metal concentration seem to have an effect on this aging process. Aging for up to 2 yr at 23°C induced the transformation of an initially noncrystalline alumina to more ordered products including gibbsite. Results indicated Cu movement toward the surface of the coprecipitate with increasing aging time was observed, which may have been caused by the transformation of noncrystalline alumina to more ordered products. Prolonged aging resulted in decreased Cu solubility after coprecipitate formation, with the percent of copper retained decreasing as the initial Cu concentration increased (Martinez and McBride 2000). The reversibility of metal partitioning on ferrihydrite could also be significantly influenced by the transformation to more thermodynamically stable structures such as goethite or hematite. During a 400-hr ferrihydrite transformation period, the extent of sorption increased and sorption reversibility decreased significantly for Mn and Ni. For both Pb and Cd, a net desorption with aging was demonstrated and sorption reversibility remained essentially unchanged. These differences in metal behavior were consistent with structural incorporation of Mn and Ni into the goethite or hematite structure and minimal incorporation of Cd and Pb within these crystalline products at pH 6 (Ford et al. 1997). Laboratory-synthesized ferrihydrite showed the most distinct transformation to goethite and hematite during an equilibration period of 60 d at 70°C. Ferrihydrite transformation to goethite and hematite resulted in increased Pb in solution, with free Pb activity about two orders of magnitude higher after heating the suspensions. Lead activity increased after heating soils with a high metal oxyhydroxide content, which, however, may be partly explained by an increase of dissolved organic carbon (DOC) after thermal treatment (Martinez et al. 1999). Aging freshly precipitated hydrous aluminum and iron oxides up to 12 hr markedly reduced boron adsorption; beyond 12 hr the effect of aging diminished (McPhail et al. 1972). Boron retention by hydroxy iron and aluminum materials was significantly reduced by aging on a steam bath before being treated with B, with differences being larger for the materials aged for longer periods (Sims and Bingham 1968). The Zn adsorption capacities were the same for the aged Fe oxide (goethite) and the aged Al oxide (gibbsite). The adsorption capacities for both fresh oxides (amorphous) were about 10 times those for the aged oxides, which corresponded to a 10-fold difference between their respective cationexchange capacities (CECs) and surface areas (Shuman 1977). From these data, it can be concluded that transformation of metal oxyhydroxides decreases metal availability. However, although these transformations can be a relatively rapid process under laboratory conditions (freshly prepared metal oxyhydroxides, elevated temperatures, a high liquid to soil ratio, etc.), these transformations usually occur at a much slower rate in natural soils and the influence of metal oxyhydroxides on metal availability will therefore be less important in natural soils compared to the batch experiments with freshly prepared metal oxyhydroxides. The sorption of cadmium to calcite in aqueous solutions is characterized by two steps; the first reaches completion within 24 hr while the second step proceeds at a slow and nearly constant rate for at least 7 d (Davis et al. 1987). It

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was proposed that the second step was due to crystal growth during recrystallization of calcite. As the precipitate ages, the rate of recrystallization will eventually decrease because the number of surface defects and small crystals will decrease to minimize the free energy of the surface. Also, Ni was found to coprecipitate with calcite as a result of recrystallization during a 42-d aging period (Carlsson and Aalto 1998). For the same reasons as for metal oxyhydroxides, batch experiments with freshly prepared calcite are not representative for aging processes occurring in natural soils. Samples of bentonite, illite, and kaolinite clays were studied for their Zn fixation properties under various pH levels, alternate wetting and drying conditions, and incubation at moisture saturation. Bentonite and illite fixed (not removed by 1 N NH4OAc) Zn during repeated wetting and drying. Zn fixation was directly related to pH and the amount of Zn added. Incubation of treated samples at soil moisture saturation resulted in approximately half the amount of Zn fixed when subjected to repeated wetting and drying. Compared to bentonite and illite, kaolinite fixed relatively small amounts of Zn, regardless of the treatment (Reddy and Perkins 1974). With increasing residence times, up to 1 yr, all precipitate phases with talc and pyrophyllite drastically increased in stability, as was documented by decreasing amounts of Ni released by HNO3 and EDTA treatments (Scheckel et al. 2000). The dissolved concentrations of Ni and Zn in presence of Al montmorillonite and Al-13 montmorillonite decreased significantly over 30 wk (Lothenbach et al. 1997, 1998, 1999). In contrast, the large Cd cations were partially remobilized by excess Ba. The large cation Pb seemed to be bound specifically to montmorillonite regardless of its size. Cd and Pb immobilization did not change within 60 wk. The Al3+ ions preferentially resided in octahedral sites, although in many precipitates tetrahedral sites were also occupied. Some of these Al3+ cations can be replaced by other cations with similar ionic radius (isomorphous substitution), e.g., Ni, Zn, and Cu, whereas the incorporation of the larger cations Cd and Pb would disturb this structure. The aging process of the polynuclear aluminum complex Al-13 had no effect on the dissolved metal concentrations or on the remobilization of metals by Ba. Metal sorption on pure montmorillonite was not affected by aging (Lothenbach et al. 1997, 1998). After addition of 10 mM HCl to the suspension, a positive correlation was observed between the amount of remobilized aluminum and dissolved zinc in the presence of Al montmorillonite; this indicates that Zn was incorporated in the aluminum lattice and remobilized only after the dissolution of the aluminum hydroxide lattice. A slope of nearly 1 between dissolved Zn and Al indicated a rather uniform distribution of the incorporated Zn in the relatively amorphous Al hydroxide layers of fresh Al montmorillonite. In contrast to Zn, Cd sorbed on fresh and aged Al montmorillonite was remobilized before a substantial amount of Al was released into solution. This preferential release indicated that Cd was sorbed on surface sites of the Al hydroxide lattice and that no or only little incorporation occurred (Lothenbach et al. 1999). The 54-d Cd adsorption and desorption isotherms overlapped, which implied that for illite the sorption process is com-

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pletely reversible (Comans 1987). Lattice penetration strongly depended on the mineral system used, the amount of Co and Zn added, and, in particular, the pH and the reaction time. These specific sorption reactions did not readily attain equilibrium but after an initial period of several days tended to approach a nearly steady-state reaction. The variation in reactivity between isomorphous series was, in general, similar to that known for the resistance of minerals to acid attack and to natural weathering processes. The 2:1 layer minerals were more reactive than the 1:1 layer minerals. Within each group, the reactivity was again related to stability to weathering and acid attack (Tiller and Hodgson 1960). The desorption rate and the total quantity of Ni released from the kaolinite surface was greater for the short-term than for the long-term experiments (Eick et al. 2001). The amount of Zn fixed by kaolinite had been found earlier to increase with grinding, which may be attributed to newly exposed surfaces (Elgabaly 1944). Also, the amount of Zn sorbed in Aiken clay loam increased as a result of grinding because grinding tends to break down clay minerals, which results in increased surface where fixation can take place (Brown 1950). The aforementioned data clearly indicate that in batch experiments with clays metal availability usually decreases with time, with the influence of aging depending on the metal. Furthermore, several environmental parameters such as drying and rewetting cycles and pH are reported to have an effect on this aging process. As the batch experiments were mostly carried out with freshly ground clays, it can be expected that, in these studies, the influence of clay on metal aging is overestimated compared to natural soils where clays are weathered and have less exposed surfaces. Total dissolved Cd, Cu, Pb, and Zn increased in an organic system after 200 d of aging, and a parallel increase in DOC concentration was observed (Martinez and McBride 1999). With time, organic matter breaks down into smaller particles, which results in an increase of the DOC concentration, leading to raising dissolved metal concentrations. B. Natural Soils It can be concluded from the previous section that experiments with soil components are not representative of what is happening in the field. In this section, experiments with natural soils are discussed as these are better suited to assess the effect of aging on metal availability in the field. After 1 wk of aging, the Cu, Zn, Mn, and Fe extracted from 11 spiked Colorado soils decreased. Of all metals, Cu remained most available, which is in agreement with the long-term residual effect often reported for Cu fertilizers (Follett and Lindsay 1971). The observed decline in Cd, Cu, Ni, Pb, and Zn content, over time after spiking, when extracted with H2O2-NH4OAc was faster in samples collected from a warm environment (25°C) than those from cooler conditions (15°C) (Hooda and Alloway 1994). The concentration of Cd, Ni, Co, and Zn in solution of a yellowish-brown loamy sand from Western Australia decreased after addition to the soil (up to 30 d). For Cd, the decreases were slow; for Zn, and especially Ni

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and Co, these were faster, which might have been related to the larger ionic radius of Cd (Barrow 1998). Xiang and Banin (1996) investigated the transformations of native and added Cd, Cr, Cu, Ni, and Zn in two arid zone soils during 1 yr of incubation at saturated past water regimen. Metals added in a soluble form were slowly repartitioned according to the initial pattern of distribution in the given soil; in other words, metal partitioning tended to return to the fraction quasiequilibrium state characteristic of the native, nonamended soil. Furthermore, Xiang and Banin (1996) observed an initial (<1 hr) fast and subsequent slow reduction of lability and addition of higher metal concentrations resulted in higher labile fractions. Conversion of Zn to forms not extracted with 0.1 M HCl from a fine sandy loam soil from Washington was a gradual process involving a period of several years (Boawn et al. 1960). Dithizon-extracted (diphenylthiocarbazone) Zn in six USA soils decreased gradually with time (Brown et al. 1964). After 3 d, the 65 Zn in solution of 25 agricultural soils continued to decrease slowly (Tiller et al. 1972). The amount of Zn desorbed from six California rice soils by diethylenetriaminepentaacetic acid (DTPA) continued to decrease with increasing aging time. Elevated temperature further enhanced Zn aging and reduced Zn extractability. The reduction of Zn extractability was rapid during the first few weeks of aging and declined slowly afterward (Kuo and Mikkelsen 1980). Barrow (1986) incubated soil samples with Zn nitrate solutions from 1 to 30 d at temperatures from 4° to 60°C. Soluble Zn concentrations decreased with increasing period of incubation, with the highest rate of decrease at high incubation temperatures. The change with time was highest at low levels of Zn addition. The effects of both temperature and time were closely described by a model that postulated an initial rapid adsorption of ZnOH+ ions onto heterogeneous charged surfaces, followed by diffusive penetration. The results showed that the slow reaction between soil and added Zn continued for a very long period: the reaction appeared to be continuing even at the longest period at the highest temperature used, which was calculated to be equivalent to about 500 d at 25°C (Barrow 1986). The extraction of soil Zn by DTPA during a 30-d incubation period demonstrated a consistent decrease in the availability of applied Zn (Brennan and Gartrell 1990). DTPA-extracted Zn (pH 8.5, 64% clay) decreased with incubation times up to 180 d in a calcareous soil from Victoria. Higher temperatures and repeated drying and rewetting of the soils increased the fixation rate (Ma and Uren 1997b). The proportions of recently added Zn in the EDTA-extracted fraction of some calcareous soils also decreased with increasing time and temperature, and the immobilization of Zn was further enhanced by drying and rewetting, especially at high temperatures (22° and 40°C) (Ma and Uren 1997a). Xiang et al. (1995) incubated 18 Chinese Zn-spiked soils for 7 mon and observed a transformation of the more available fractions into the more stable fractions; more of such transformation occurred in calcareous soils than in acid and neutral soils. Sorption of Zn in a Chinese calcareous soil increased between 10 and 1000 min after spiking, and the adsorption rate increased with elevated temperature and pH

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(Ma and Liu 1997a). Different Zn fractions were assessed using freshly spiked artificial soils and spiked soils that had been aged for 8 wk. Standard artificial Organization for Economic Cooperation and Development (OECD) soils (70% sand, 20% kaolinite clay, and 10% Sphagnum peat adjusted to pH 6 with CaCO3) were aged in four different ways: (1) storing at 20°C, (2) percolation followed by storing at 20°C, (3) alternately heating at 60°C and storing at 20°C, and (4) alternately freezing at −20°C and storing at 20°C. Aging had no influence on the pore water concentration or the water-soluble and the calcium chloride-extracted fraction of Zn in the artificial soils (Lock and Janssen 2002). In experiments with two Danish soils (loamy sand and sandy loam) spiked with low Cd concentrations, equilibrium was reached in 1 hr, and exposures up to 67 wk did not reveal any long-term changes in Cd sorption or desorption capacities (Christensen 1984a, b). Incubation (up to 8 wk) also had no effect on the quantity of Cd removed from 13 soils of the southeastern U.S. by the extractants (King 1988). Hamon et al. (1998) found that Cd was fixed at a rate of 1%–1.5% of the total added Cd per year after fertilization with Cd-containing superphosphate. Radiolabile cadmium applied to two topsoils gradually decreased from about 100% of the adsorbed Cd at time 0 to around 35% after 500 d (Nakhone and Young 1993). 109Cd and 65Zn concentrations decreased in the mobile fractions with time. Elevated temperature reduced the mobile fraction. A slow but steady diffusion of the spiked metals toward inert sites in the soil was observed with time (1 yr) (Alma˚s et al. 1999). The addition of organic matter reduced the rate of 109Cd and 65Zn transfer into the irreversibly sorbed fractions, whereas the transfer rate increased with increasing temperature (Alma˚s et al. 2000). In contrast to Cd, sorption reactions between Zn or Ni and a suspension of the fraction smaller than 2 µm of seven soils were not completed within 14 d (Tiller et al. 1984). There was a continuing loss of Hg from solution as a function of time. Complete equilibrium between five soils and the Hg solution was not attained within 336 hr (Amacher et al. 1990). The fraction of applied Cu extracted from soils by EDTA decreased during 44 wk of incubation under both moist and dry conditions, and this decrease was larger in moist compared to dry soils (Williams and McLaren 1982). The amount of Cu desorbed declined with increasing contact time (up to 12 wk) between the added Cu and the soil (Hogg et al. 1993). Native soil Cu also showed a decrease in desorption with increasing time of initial equilibration. Two explanations were proposed: Cu was continually cycled in the soil through decaying organic materials, and a pool of Cu ions may have been maintained that did not have time to react irreversibly with the soil. Alternatively, the nature of the equilibration process in Ca(NO3)2 at a high solution/soil ratio created conditions favorable for the slow reactions that occur in soils at a normal soil moisture content and was probably much slower than those observed in soil suspensions in the laboratory. The water-extracted concentrations indicated that the same or lower Cu concentrations were extracted in spiked soils than in a field soil that had been contaminated 70 yr ago (Scott-Fordsmand et al. 2000a,b). The data for metal aging in natural soils indicate that the availability of metal ions with a small ionic radius

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such as Cu, Ni, and Zn decreases with time whereas the availability of ions with a higher ionic radius such as Cd and Pb changes little with time. Again, environmental variables such as pH, temperature, and moisture content seem to affect the rate and extent of metal fixation. Aging, however, does not always result in decreases in metal availability but may also lead to increases in metal availability if metals are originally associated with organic phases that are subject to degradation in the long term. For example, increasing incubation time decreased the availability of added inorganic Cd but increased the availability of Cd applied with sewage sludge (Street et al. 1978). However, apart from an initial enhancement in metal availability during the first year or two following application of biosolid, long-term increases in metal availability resulting from decomposition of biosolid organic matter have not been observed to date (Chang et al. 1997). Increases in EDTA-extracted metals (Pb, Cu, Cd, Zn, Ni) in response to additions of sewage sludge occurred within the first 10 yr after sludge addition (McGrath and Cegarra 1992).

III. Environmental Bioavailability In the previous section, the influence of aging on environmental availability of metals was discussed. Metal availability was assessed as the fraction present in the pore water or some kind of extract. However, this fraction does not necessarily represent the fraction that is available for an organism. It seems, therefore, more straightforward to measure the metal concentration in an organism, which should be a better measure of the fraction of a metal that is available for that particular organism. A. Plants Zn concentrations in sweet corn (Zea mays), sampled on a calcareous and a noncalcareous Washington soil with similar pH, decreased during a 5-yr period. This decrease corresponded to a decrease in 0.005 M DTPA- and 0.1 N HClextracted Zn concentrations. Extracted Zn in the calcareous soil declined more rapidly compared to the noncalcareous soil; however, as the conversion of Zn to nonextractable forms approached an equilibrium, similar fractions of Zn were extracted (Boawn 1974). Incubation for 30 d at 30°C of 54 Australian soils spiked with Zn decreased the concentration of Zn in subterranean clover shoots as well as DTPA-extracted Zn for all soils relative to freshly spiked soils (Brennan 1990). In a 13-yr experiment with a Cu-spiked yellow-brown lateritic sandy earth in a semiarid mediterranean environment, the internal Cu concentration in wheat plants declined linearly with time (Brennan et al. 1986). Bush bean (Phaseolus vulgaris) was grown in plots spiked with Cd, Tl, and V after 1, 18, and 30 mon of aging. All plants from the 1-mon planting died within 10 d of germination. No significant decrease in the metal content of the plants was observed between 18 and 30 mon of aging. These results are in agreement with the 0.05 M HCl–0.05 M H2SO4 extractions, which showed that after 20 min

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extracted metal concentrations were significantly higher in comparison with the extracts 18 and 30 mon after treatment. No detectable decrease was observed during the last 12 mon (Martin and Kaplan 1998). Aging of 10 Cd-spiked soils for 1–14 d did not substantially reduce Cd availability or accessibility for isotopic exchange. Varying the soil incubation procedure after soil spiking and before plant growth only marginally affected the specific activity of Cd in plants (Smolders et al. 1999). In some cases, metal bioavailability can also increase with time, as was shown for Cd and Pb uptake by ryegrass 1 yr after sludge-amended soils had been treated with metal salts (Hooda and Alloway 1993). Such an increase in bioavailability could result from breakdown of organic matter or a reduced pH. From this review, it can be concluded that metal availability estimated by some kind of extraction procedure is usually correlated to the internal concentration in plants. Environmental availability, expressed as pore water metal concentrations or the metal fraction extracted with a salt solution, therefore seems to be a good predictor of environmental bioavailability to plants. B. Invertebrates Invertebrates are often able to regulate the internal concentration of essential metals within a narrow range. Therefore, it is not surprising that no relationship was found between the internal Cu concentration in the earthworm (Eisenia fetida), the chronic Cu toxicity to E. fetida, and the water-extracted Cu after exposure to freshly spiked and contaminated field soils (Scott-Fordsmand et al. 2000b).

IV. Toxicological Bioavailability In the section on environmental availability, the influence of aging on metal accumulation was discussed. However, the effect of a metal on an organism is not always proportional to the metal concentration in that species. The only way to assess the influence of aging on the effect caused by metal is, therefore, to evaluate metal toxicity using bioassays. A. Plants The ecotoxicity of Zn to red clover (Trifolium pratense) was evaluated in soils originating from a gradient of metal concentrations near a Zn smelting factory in Budel, the Netherlands. Toxicity experiments with soils adjusted to a pH of 5.2 with CaCO3 resulted in 24-d EC50s (wet weight) of 347 mg Zn/kg dry wt and 215 mg Zn/kg dry wt for the shoots and the roots, respectively. In addition, an uncontaminated soil with a background concentration of 8 mg Zn/kg dry wt originating from 15 km away from the factory, adjusted to pH 5.2, was spiked with 32, 100, 320, and 1000 mg Zn/kg dry wt. EC50s at 24 d of 131 and 68 mg Zn/kg dry wt were determined for shoot and root wet weight, respectively. Although the 24-d EC50 (total Zn) in the spiked soils and the contaminated field

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soils differed by a factor of more than two, the EC50 calculated on the basis of the 0.01 M CaCl2-extracted fraction differed less than 25% (van der Hoeven and Henzen 1994a,b). The dry weight of shoots of Phaseolus vulgaris in Zn-deficient soils with recently applied Zn was consistently higher than in soils that were incubated for 15 d at 40°C. The recovery of added Zn in four soils from Western Australia and Queensland decreased exponentially in all extractants with increasing time (up to 8 d at 40°C) of incubation in all soils (Armour et al. 1989). On 54 Zn-deficient Australian soils, where growth of subterranean clover was increased by Zn application, prior incubation for 30 d of the spiked soils led to decreased plant growth. This growth reduction was reflected in a decrease in DTPA-extracted Zn after incubation (Brennan 1990). Addition of Cu to Cu-deficient soils collected in Western Australia resulted in a significant yield increase in wheat plants (Brennan et al. 1980). This increase in yield, however, significantly decreased with incubation time. Also, ammonium oxalate-extracted Cu concentrations decreased during the 120-d aging period. In a subsequent study, it was shown that higher temperature and longer incubation time decreased the dry weight of wheat grown on a black calcareous sand (Brennan et al. 1984). From these data, it can be concluded that toxicity as well as deficiency in plants are related to the chemical availability of metals. B. Invertebrates A few studies have compared the toxicity of Zn spiked soils with the toxicity in historically contaminated field soils in the vicinity of a Zn smelter. The toxicity of Zn to the earthworm E. fetida in freshly spiked artificial OECD soil was at least 10-fold higher than in soils collected around the Avonmouth (UK) smelter (Spurgeon and Hopkin 1995). The differences in Zn toxicity between these sets of soils may be related to difference in aging, forms of added Zn, and soil properties. Smit and Van Gestel (1996) compared the difference in Zn toxicity to springtails between soils collected around the Budel smelter (the Netherlands) and the control soil to whereas Zn salts were added at concentrations that are similar to those of the smelter soils. Soil pH of the smelter soils was adjusted and varied from 5.5 to 6.2 whereas the pH values of the spiked soils were 5.3 to 5.5. The EC50 for springtail reproduction in the spiked soils was 210 mg Zn/ kg dry wt whereas there was no effect on reproduction in the smelter soils up to the highest concentration of about 1500 mg Zn/kg dry wt. This result indicates that toxicity data obtained with freshly spiked soils overestimate effects found in field contaminated soils. For enchytraeids, Zn toxicity in freshly spiked artificial soil was similar to that in field soils sampled near a Zn smelter (Posthuma and Notenboom 1996); however, soil pH in the field soils was approximately one unit lower than in the artificial soil. In two other studies, Smit et al. (1997) and Smit and Van Gestel (1998) compared Zn toxicity to the springtail Folsomia candida immediately after spiking and after aging under field and laboratory conditions for up to 26 mon. In addition to this study with F. candida, Posthuma et al. (1998) described a

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similar study using the same soils with the oligochaete Enchytraeus crypticus. For both organisms, Zn toxicity decreased considerably after aging, but a concurrent increase in pH of at least one unit was observed. For the same soils, the total number of nematodes and the number of nematode taxa were determined up to 22 mon after treatment. It was found that the initial effect of Zn spiking gradually decreased with time, suggesting recovery with aging. Also in this case, the concomitant increase of soil pH during the aging period may have affected the recovery process. Therefore, it remains unclear to what extent the observed reduced Zn toxicity for these studies was due to the pH changes or to the effect of aging. Recently, Lock and Janssen (2002) exposed Enchytraeus albidus to standard artificial soil spiked with Zn and assessed the influence of aging on Zn toxicity. Acute and chronic toxicity were similar in freshly spiked artificial soils and soils aged for 2 mon, which agrees with the lack of change in chemical availability in these artificial soils. This result indicates that aging had little or no effect on Zn toxicity and availability in standard artificial soils. Using a central composite design, Lock and Janssen (2001b) developed a model predicting chronic Zn toxicity to F. candida in freshly spiked artificial soils as a function of pH, cation-exchange capacity, and total Zn concentration. However, application of this model to toxicity tests performed with 19 Zncontaminated field soils resulted in overpredictions, indicating that aging decreased Zn toxicity (Lock and Janssen 2001b). No differences in the 0.01 M CaCl2-extracted Cu concentration and Cu toxicity to Folsomia fimetaria were found between 1 d and 12 wk of aging. However, in a historically contaminated field soil with similar soil characteristics, less Cu was extracted, and reproduction of F. fimetaria was less affected at comparable total Cu concentrations (Pedersen and van Gestel 2001). A 21-d EC10 of 337 (207–581) mg Cu/kg dry wt and a 21-d EC50 of 994 (591–1208) mg Cu/kg dry wt were observed for Folsomia fimetaria exposed in soils spiked with Cu 1 d before the experiment. Using soil from a field site contaminated with Cu more than 70 yr previously, no effect was observed at concentrations as high as 2911 mg Cu/kg dry wt (Scott-Fordsmand et al. 2000a). Also for Eisenia fetida, the 21-d EC10 for cocoon production was 34 (8–181) mg Cu/kg dry wt and the 21-d EC50 was 210 (152–258) mg Cu/kg dry wt for soils spiked 1 d before the experiment, whereas a 21-d EC10 of 248 (134–317) mg Cu/kg dry wt and a 21-d EC50 of 517 (330– 617) mg Cu/kg dry wt were found for the field soils contaminated 70 yr previously (Scott-Fordsmand et al. 2000b). The data on the toxicological bioavailability of metals seem to agree with chemical availabilility. However, differences in metal toxicity between aged and freshly spiked soils were difficult to assess because most studies were confounded by differences in soil properties. It should also be mentioned that if some species have mechanisms to facilitate desorption or to overcome the binding of sequestered metals, the amount of the metal that will be available or the rate of its uptake will differ from organisms not having such mechanisms (Alexander 2000). Should it be found that accessibility varies among species or even

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by route of exposure (e.g., dermal, intestinal, respiratory), a single assay of bioavailability will not be sufficient. C. Microorganisms Unlike studies with invertebrates and plants, microbial assays involve the mass or activity of indigenous soil organisms. An important consequence is that metal-spiked soil cannot be aged in the absence of the microorganisms. Furthermore, microorganisms will react completely differently to sudden large inputs of soluble metals compared to a slow increase of the metal concentration, which is characteristic in field situations. When considering metal tolerance, it is also important to distinguish between the development of new tolerance in a population in response to metal toxicity and the selection of individuals within a population that are intrinsically more tolerant to metals (Giller et al. 1998). In addition, effects of Zn toxicity on microbial populations may only become apparent after months or even years of exposure. For example, Rhizobium populations in Zn-spiked soils showed no response after 2, 6, and 12 mon, but after 18 mon the population decreased at higher Zn concentrations (Chaudri et al. 1992). However, during these long time periods, metal aging also is taking place. Despite their sensitivity to metal toxicity, microbial toxicity assays on indigenous microorganisms do not seem to be very useful for studying the effect of aging on metal availability for all the reasons just discussed.

V. Conclusions and Recommendations An abundance of data support the theory that metal availability often decreases with time because of aging processes. However, it is still difficult to assess the rate and extent of metal aging because several environmental parameters seem to affect the aging processes simultaneously, such as temperature (Almas et al. 1999, 2000; Barrow 1986; Barrow et al. 1989; Bruemmer et al. 1988; Gerth et al. 1993; Hooda and Alloway 1994; Kuo and Mikkelsen 1980; Ma and Liu 1997; Ma and Uren 1997a,b; Trivedi and Axe 2000), repeated drying and rewetting (Ma and Uren 1997a,b; Reddy and Perkins 1974), soil moisture content (Hogg et al. 1993; Williams and McLaren 1982), pH (Barrow et al. 1989; Brennan 1990; Bruemmer et al. 1988; Ma and Liu 1997; Reddy and Perkins 1974; Schultz et al. 1987; Tiller and Hodgson 1960; Xiang et al. 1995), and total metal concentration (Barrow 1986; Eick et al. 2001; Martinez and McBride 2000; Nelson and Melsted 1955; Reddy and Perkins 1974; Xiang and Banin 1996). Temperature, drying and rewetting, soil moisture content, and pH may all have an influence on the rate of the aging process, but pH seems to be the major factor determining the extent of the influence of aging on metal availability. Although pH and the amount of available sorption sites, which is related to the cation-exchange capacity (CEC), are the main factors determining the initial adsorption of metals to the soil (Anderson and Christensen 1988; Buchter et al.

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1989; Harter 1982; Lee et al. 1996; Sauve´ et al. 2000) and the toxicity of freshly metal-spiked soils (Crommentuijn et al. 1997; Lock et al. 2000; Lock and Janssen 2001c,d; Spurgeon and Hopkin 1996), pH is probably the most important parameter influencing the aging process. It should be possible to estimate the influence of aging on metal availability for a given soil by calculating the difference between the metal fraction sorbed immediately after spiking and the fraction that is predicted to be sorbed after aging on the basis of the pH. For example, in acidic soils, the sorbed metal fraction will remain quite constant over time after spiking because metals stay in solution at a low pH. In contrast, in soils with a high pH, the influence of aging on metal availability will be high if the initial sorption is low but smaller when the initial sorption is already high. This difference occurs because the metal fraction sorbed during aging is high at a high pH, independent of the initial sorption. To illustrate that metal partitioning in aged soils is related to pH, partition coefficients (Kd, L/kg) for Cd, Cr, Cu, Ni, Pb, and Zn of 21 uncontaminated and 25 historically contaminated field soils were calculated on the basis of the data reported by De Groot et al. (1998). Partitioning coefficients were calculated according to the formula Kd = [M]S/ [M]W, with [M]S the total soil concentration (mg/kg) and [M]W the pore water concentration (mg/L). Indeed, the calculated partition coefficients are found to be related to soil pH, and this relationship was similar for uncontaminated and contaminated field soils (Fig. 1). To work out the foregoing hypothesis in more detail, it would be interesting to collect contaminated and uncontaminated field soils with the same soil properties and subsequently spike the uncontaminated soils with the same metal concentrations as the contaminated soils. In this way, it would be possible to quantify the effect of aging and to monitor metal availability as a function of time. It is suggested that environmental availability, environmental bioavailability, and toxicological bioavailability, are assessed simultaneously. With this approach, it will be clear whether the results on environmental availability can be extrapolated for use in environmental risk assessment of contaminated soils.

Summary An overview is given on the literature concerning the effect of long-term processes, called aging, on metal availability, and consequences for risk assessment of metal-contaminated soils are discussed. Experiments with freshly prepared metal oxyhydroxides and calcites overpredict the influence of aging because of the initial transformations occurring in these systems. Also, freshly ground clays are not representative of field soils because the surface area exposed for fixation is unrealistically high. Experiments with field soils confirm the hypothesis that metal availability can decrease as a result of aging, especially for metals with a small ionic radius. Although aging also seems to affect metal uptake, it should be noted that some organisms are able to regulate their internal metal concentrations within a narrow range, especially for essential elements. Deficiency of essential metals as well as metal toxicity to plants and invertebrates seem to be

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Fig. 1. Partition coefficients (Kd) of cadmium (A), chromium (B), copper (C), nickel (D), lead (E), and zinc (F) in function of pH(CaCl2) in uncontaminated (full circles) and contaminated, (open circles) soils. (Calculated from De Groot et al. 1998.)

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related to the environmental availability of these metals. Unfortunately, aging effects are often confounded by differences in soil properties between freshly spiked and historically contaminated soils. Environmental parameters such as temperature, moisture content, drying and rewetting cycles, and pH affect the rate of aging, whereas pH seems to be the sole important parameter affecting the extent of aging.

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clear aluminium and montmorillonite compounds. Environ Sci Technol 31:1452– 1462. Lothenbach B, Krebs R, Furrer G, Gupta SK, Schulin R (1998) Immobilization of cadmium and zinc in soil by Al-montmorillonite and gravel sludge. Eur J Soil Sci 49: 141–148. Lothenbach B, Furrer G, Scha¨rli H, Schulin R (1999) Immobilization of zinc and cadmium by montmorillonite compounds: effect of aging and subsequent acidification. Environ Sci Technol 33:2945–2952. Ma YB, Liu JF (1997) Adsorption kinetics of zinc in a calcareous soil as affected by pH and temperature. Commun Soil Sci Plant Anal 28:1117–1126. Ma YB, Uren NC (1997a) The effects of temperature, time and cycles of drying and rewetting on the extractability of zinc added to a calcareous soil. Geoderma 75:89–97. Ma YB, Uren NC (1997b) The fate and transformations of zinc added to soils. Aust J Soil Res 35:727–738. Martin HW, Kaplan DI (1998) Temporal changes in cadmium, thallium and vanadium mobility in soil and phytoavailability under field conditions. Water Air Soil Pollut 101:399–410. Martinez CE, McBride MB (1998) Coprecipitates of Cd, Cu, Pb and Zn in iron oxides: solid phase transformation and metal solubility after aging and thermal treatment. Clays Clay Miner 46:537–545. Martinez CE, McBride MB (1999) Dissolved and labile concentrations of Cd, Cu, Pb, and Zn in aged ferrihydrite-organic matter systems. Environ Sci Technol 33:745–750. Martinez CE, McBride MB (2000) Aging of coprecipitated Cu in alumina: changes in structural location, chemical form, and solubility. Geochim Cosmochim Acta 64: 1729–1736. Martinez CE, Sauve´ S, Jacobson A, McBride MB (1999) Thermally induced release of adsorbed Pb upon aging ferrihydrite and soil oxides. Environ Sci Technol 33: 2016–2020. McBride M, Martinez CE, Sauve´ S (1998) Copper(II) activity in aged suspensions of goethite and organic matter. Soil Sci Soc Am J 62:1542–1548. McGrath SP, Cegarra J (1992) Chemical extractability of heavy metals during and after long-term applications of sewage sludge to soil. J Soil Sci 43:313–321. McPhail M, Page AL, Bingham FT (1972) Adsorption interactions of monosilicic and boric acid on hydrous oxides of iron and aluminum. Soil Sci Soc Am Proc 36:510– 514. Nakhone LN, Young SD (1993) The significance of (radio-) labile cadmium pools in soil. Environ Pollut 82:73–77. Nelson JL, Melsted SW (1955) The chemistry of zinc added to soils and clays. Soil Sci Soc Am Proc X:439–443. Padmanabham M (1983) Adsorption-desorption behaviour of copper (II) at the goethitesolution interface. Aust J Soil Res 21:309–320. Pedersen MB, van Gestel CAM (2001) Toxicity of copper to the collembolan Folsomia fimetaria in relation to the age of soil contamination. Ecotoxicol Environ Saf 49: 54–59. Posthuma L, Notenboom J (1996) Toxic effects of heavy metals in three worm species exposed in artificially contaminated soil substrates and contaminated field soils. RIVM-report no. 719102048, Bilthoven. Posthuma L, Van Gestel CAM, Smit CE, Bakker DJ, Vonk JW (1998) Validation of toxicity data and risk limits for soils: final report. National Institute of Public Health

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and the Environment (RIVM). RIVM report 607505004. Bilthoven, The Netherlands. Reddy MR, Perkins HF (1974) Fixation of zinc by clay minerals. Soil Sci Soc Am Proc 38:229–231. Sauve´ S, Hendershot W, Allen HE (2000) Solid-solution partitioning of metals in contaminated soils: dependence on pH, total metal burden, and organic matter. Environ Sci Technol 34:1125–1131. Scheckel KG, Sparks DL (2000) Kinetics of the formation and dissolution of Ni precipitates in a gibbsite/amorphous silica mixture. J Colloid Interface Sci 229:222–229. Scheckel KG, Scheinost AC, Ford RG, Sparks DL (2000) Stability of layered Ni hydroxide surface precipitates—a dissolution kinetics study. Geochim Cosmochim Acta 64: 2727–2735. Schultz MF, Benjamin MM, Ferguson JF (1987) Adsorption and desorption of metals on ferrihydrite: reversibility of the reaction and sorption properties of the regenerated solid. Environ Sci Technol 21:863–869. Scott-Fordsmand JJ, Henning Krogh P, Weeks JM (2000a) Responses of Folsomia fimetaria (Collembola: Isotomidae) to copper under different soil copper contamination histories in relation to risk assessment. Environ Toxicol Chem 19:1297–1303. Scott-Fordsmand JJ, Weeks JM, Hopkin SP (2000b) Importance of contamination history for understanding toxicity of copper to earthworm Eisenia fetida (Oligochaeta: Annelida), using neutral-red retention assay. Environ Toxicol Chem 19:1774–1780. Shuman LM (1977) Adsorption of Zn by Fe and Al hydrous oxides as influenced by aging and pH. Soil Sci Soc Am J 41:703–706. Sims JR, Bingham FT (1968) Retention of boron by layer silicates, sesquioxides, and soil materials: II. Sesquioxides. Soil Sci Soc Am Proc 32:364–369. Smit CE, van Gestel CAM (1996) Comparison of the toxicity of zinc for the springtail Folsomia candida in artificially contaminated and polluted field soils. Appl Soil Ecol 3:127–136. Smit CE, van Bellen P, van Gestel CAM (1997) Development of zinc bioavailability and toxicity for the springtail Folsomia candida in an experimentally contaminated field plot. Environ Pollut 98:73–80. Smit CE, van Gestel CAM (1998) Effects of soil type, prepercolation, and aging on bioaccumulation and toxicity of zinc for the springtail Folsomia candida. Environ Toxicol Chem 17:1132–1141. Smolders E, Brans K, Fo¨ldi A, Merckx R (1999) Cadmium fixation in soils measured by isotopic dilution. Soil Sci Soc Am J 63:78–85. Spurgeon DJ, Hopkin SP (1995) Extrapolation of the laboratory-based OECD earthworm toxicity test to metal-contaminated field sites. Ecotoxicology 4:190–205. Spurgeon DJ, Hopkin SP (1996) Effects of variations of the organic matter content and pH of soils on the availability and toxicity of zinc to the earthworm Eisenia fetida. Pedobiologia 40:80–96. Street JJ, Sabey , Lindsay WL (1978) Influence of pH, phosphorus, cadmium, sewage sludge, and incubation time on the solubility and plant uptake of cadmium. J Environ Qual 7:286–290. Tiller KG, Hodgson JF (1960) The specific sorption of cobalt and zinc by layer silicates. Clays Clay Miner 9:393–403. Tiller KG, Honeysett JL, De Vries MPC (1972) Soil zinc and its uptake by plants. I. Isotopic exchange equilibria and the application of tracer techniques. Aust J Soil Res 10:151–164. Tiller KG, Gerth J, Bru¨mmer G (1984) The sorption of Cd, Zn and Ni by soil clay

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fractions: procedures for partition of bound forms and their interpretation. Geoderma 34:1–16. Trivedi P, Axe L (2000) Modeling Cd and Zn sorption to hydrous metal oxides. Environ Sci Technol 34:2215–2223. van der Hoeven N, Henzen L (1994a) Groei van de plantensoorten Lolium perenne, Vicia sativa en Trifolium pratense op grond uit Budel en effecten van zink en cadmium op die groei. Report IMW-TNO 94/004, . van der Hoeven N, Henzen L (1994b) Groei van de plantensoort Trifolium pratense op elf gronden uit Budel gemonsterd in een gradie¨nt. Report IMW-TNO 94/041, . van Gestel CAM (1997) Scientific basis for extrapolating results from soil ecotoxicity tests to field conditions and the use of bioassays. In: Van Straalen NM, Løkke H (eds) Ecological Risk Assessment of Contaminants in Soil. Chapman & Hall, London, pp 25–50. Van Straalen NM, Denneman CJ (1989) Ecotoxicological evaluation of soil quality criteria. Ecotoxicol Environ Saf 18:241–251. Williams JG, McLaren RG (1982) Effects of dry and moist incubation of soils on the extractability of native and applied soil copper. Plant Soil 64:215–224. Xiang HF, Banin A (1996) Long-term transformations and redistribution of potentially toxic heavy metals in arid-zone soils incubated: I. Under saturated conditions. Water Air Soil Pollut 95:399–423. Xiang HF, Tang HA, Ying QH (1995) Transformation and distribution of forms of zinc in acid, neutral and calcareous soils of China. Geoderma 66:121–135. Manuscript received September 6; accepted September 11, 2002.

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 Springer-Verlag 2003

Rev Environ Contam Toxicol 178:23–52

Algal Toxicity Tests for Environmental Risk Assessments of Metals Colin R. Janssen and Dagobert G. Heijerick Contents I. Introduction ............................................................................................................ II. Use and Misuse of Algal Toxicity Data: Regulatory Context ............................. III. Test Methods ......................................................................................................... A. General .............................................................................................................. B. Environmental Availability and Physicochemical Factors .............................. C. Biological Availability and Biological factors ................................................ IV. Adsorption, Uptake, and Toxicity Mechanisms for Metals ................................. V. Acclimation and Adaptation Under Laboratory Conditions ................................ VI. Laboratory to Field Extrapolation ......................................................................... Summary ...................................................................................................................... References ....................................................................................................................

23 25 26 26 27 34 39 41 42 42 44

I. Introduction Although algae were used as early as 1910 in toxicity tests (Allen and Nelson 1910), it was not until the 1960s that a standardized assay with freshwater algae was developed. Increased concern about the eutrophication potential of wastewaters led to the development of the Algal Assay Procedure Bottle Test (AAP) (USEPA 1971). The aim of this method was to determine the algal growth potential and algal productivity of natural waters and effluents: these tests were not specifically designed for monitoring the toxic effects of chemicals. Eventually, the AAP test, mostly using Selenastrum capricornutum (currently renamed Pseudokirchneriella subcapitata; Hindak 1990), although other species of green algae and diatoms were also recommended, was adapted for widespread use in toxicity tests and has served as a template for most standard procedures for assessing the potential toxicity of substances and wastes (Lewis 1990; Miller et al. 1978; Parrish 1985). As a result, most current regulatory methods, including the selection of test species, culture, and test methods, rely heavily on this origi-

Communicated by George W. Ware. C.R. Janssen ( ) Laboratory of Environmental Toxicology and Aquatic Ecology, Ghent University, J. Plateaustraat 22, B-9000 Ghent, Belgium. D.G. Heijerick EURAS, Rijvisschestraat 118 bus3, B-9052 Zwijnaarde, Belgium.

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nal test procedure. For example, most algal bioassays in recent years have been performed predominantly with S. capricornutum (Benenati 1990; Chiaudani and Vighi 1978; Lewis 1990), leading to justified concerns about the sensitivity of this species and its relevance for other algal taxa (Lewis 1990; Nyholm and Ka¨llqvist 1989; Wangberg and Blanck 1988). Although most standard algal assays used for regulatory purposes seem very similar in design and protocol, subtle differences in all aspects of the toxicity test design may lead to a large variability in test results. Indeed, it is unclear whether the differences in sensitivity among various algal taxa and within individual species reported in the current literature are real or are artifacts caused by differences among the various “standard” test protocols in biotic and abiotic factors. For example, the composition of the culture and test medium, which may vary considerably among test protocols, not only can determine the bioavailability of the test substance but also can affect the physiological state of the test organism before and during the bioassay and thus influence the test result. It has been shown that using starved algae as inoculum for toxicity assays can lead to higher toxicity, probably as a result of the poorer physiological condition of the algae or increased toxicant uptake. Increased nutrient concentrations, on the other hand, appear to reduce the uptake of certain substances (Ho¨rnstro¨m 1990). For example, Sanders (1979) observed a decreased uptake of arsenic when algae were exposed to higher phosphorus concentrations. Although the need for a relatively simple algal toxicity test as part of a battery of tests for regulatory toxicity testing of substances and wastes is obvious, it should be equally clear that we need to understand the meaning of the results and that we should be aware of the limitations in applying these results in environmental risk assessments (Nalewajko and Olaveson 1998). All too often, these types of data are used for regulatory purposes other than those for which they were developed. Several authors, comparing different algal toxicity test procedures, have identified potential uses and limitations of these methods and have come to similar conclusions. Lewis (1995) provides an overview of currently used toxicity tests with freshwater algae and discusses the potential implications of differences in culture techniques, nutrient media, test species and sensitivity, test concentrations, condition and duration, effects measurements, reproducibility, and use of results. Similarly, Nalewajko and Olaveson (1998) critically reviewed all aspects of algal toxicity testing and concluded that an ecophysiological approach is required to increase the ecological relevance of algal toxicity tests. They emphasized the need for increased understanding of the relationship between the physicochemical characteristics of the environment and the physiological dynamics of algae. Ho¨rnstro¨m (1990) made a survey of the different aspects of batch algal test methods that are poorly described in existing international protocols, leading to high variability in test results. Based on this review, he proposed an alternative test method that also considers prevailing environmental conditions in oligotrophic lakes. Toxicity assessment of metals and metal compounds with algae for regulatory

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hazard and risk assessment purposes is, in most cases, performed with national or international standard test procedures such as those prescribed by OECD (1984), EEC (1987), USEPA (1985), ISO (1987), and ASTM (1993). Several authors have indicated that these procedures or aspects of these procedures might not be suitable for testing metals and metal compounds; this includes the physical, chemical, and biological aspects of the test methods. Current standard methods for assessing metals suffer from a number of weaknesses from both environmental (e.g., media composition, presence of chelators, pH control) and biological availability aspects (e.g., species selection and representativeness, acclimation, adaptation). The aim of this review is to evaluate and discuss the different parameters that may affect the outcome of freshwater algal toxicity tests for metals and, where possible, to suggest solutions or modifications to the existing test protocols. This presentation should help to generate realistic metal toxicity data that can be used and incorporated in realistic risk assessment approaches and environmental management purposes. Note that the marine environment is not discussed here; such reviews have been presented by Fletcher (1991), Thursby et al. (1993), Thursby and Steele (1995), and Walsh (1982, 1988).

II. Use and Misuse of Algal Toxicity Data: Regulatory Context Laboratory algal toxicity tests may be used as screening toxicity tests, but in many cases the results are used as estimates of ecotoxic concentrations. Their use as toxicity screens is more justified than their application to predicting environmental impact (Lewis 1995). Lewis (1995) states that “the environmental relevance of results from standard laboratory algal toxicity tests is unclear.” The ASTM (1993) guideline attributes the following significance and use to the algal toxicity test: (1) to provide information on the toxicity of test materials to an important component of aquatic biota, (2) to compare the sensitivities of different species of algae and the toxicities of different materials to algae, (3) to study the effects of various environmental factors on such tests, and (4) to provide information when assessing the hazards of materials to aquatic organisms or deriving water quality criteria for aquatic organisms. Similarly, the USEPA guidelines for effluent toxicity tests with S. capricornutum suggest that “safe levels” can be established with these types of data (Weber et al. 1988, 1989). OECD (1984) and ISO (1987) state that the effects data from algal tests should not be used to predict environmental impact but only to provide the likelihood of effect or no effect. Table 1 summarizes the main test conditions of these four internationally recognized standard test procedures. The overall caution expressed in these guidelines and by several other authors (Lewis 1995; Nalewajko and Olaveson 1998) regarding appropriate application of laboratory algal toxicity data is not reflected in general regulatory practice. Most toxicity data contribute to data compilations that are then used for hazard and risk assessments. For example, the currently used European Union (EU) system for classifying chemical substances for their potential hazard to the aquatic

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Table 1. Overview of existing standard test protocols for toxicity tests with algae. Experimental variable Initial cell density (cells/mL) Test duration Initial pH Illumination (µE m−2 s−1) Temperature (°C) Number of replicates Data analysis Test acceptability

ASTM (1990)

ISO (1987)

2 × 104

1 × 104

4d 7.5–8.0

3d 8.3 (± 0.2)

30–90

120 (± 24)

24 (± 2), 20 (± 2) 3 IC50, NOEC

23 (± 2) 3; 6 in controls EC50, NOEC

USEPA (1985) 1 × 104, 7.7 × 104 4d 7.5 (± 0.2), 8.0 (± 0.1) 300 (± 25) 20 (± 2), 24 (± 1) 3 EC0, EC10, EC50, EC90

>1 × 105 Cell density 3.5 × 106 cells/mL in in controls cells/mL control should after 96 hr after 96 hr increase 16× in control after 72 hr

OECD (1984) 1 × 104 3d 8.0 120 (± 20) 21–25 (± 2) 3; 6 in controls EC50, NOEC Cell density in controls should increase 16× after 72 hr

ASTM: American Society for Testing Materials; 150, 1990; USEPA: U.S. Environmental Protection Agency; OECD: Organization for Economic Development; IC, 1984; NOEC: no observable effect concentration. Source: Adapted from Lewis (1995).

environment, under directive 67/548/EEC, uses algal toxicity test results performed according to OECD guideline 201 or the equivalent. Similarly, laboratory algal toxicity data are used in the EU risk assessment procedures for new and existing substances to derive Predicted No Effect Concentration (EC, 1996). These types of data are also used to derive water quality objectives in both the United States and Europe (Lewis 1995). Consequently it is surprising that, despite the numerous applications and the importance of this type of test, the standard protocols currently used do not adequately consider or control the various physicochemical and biological parameters that can affect the accuracy of the results. The use of some toxicity data could, therefore, have undesirable repercussions on conclusions and environmental management programs when based on erratic or nonrepresentative test results.

III. Test methods A. General Most standard algal toxicity tests are similar in technique: usually, a static test design is used to assess the potential effects of the toxicant on an algal population growing exponentially for 3–4 d in a nutrient-enriched medium (Lewis

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1995). A control and five test concentrations are used with a minimum of three replicates. Test vessels are placed on a rotary or oscillatory shaker or may be hand shaken daily. The tests are conducted under conditions of controlled temperature, light, and initial pH. Algal biomass or other endpoints determined or estimated at several intervals during the test or at the end of the assay are then used to calculate the effects of the toxicant on algae. Comparison and discussion of the general experimental conditions for widely used freshwater algal assays are given in Nyholm and Ka¨llqvist (1989), Lewis (1995), and Nalewajko and Olaveson (1998). Below, a brief comparison is given of the general experimental conditions in standard toxicity tests with algae (e.g., ASTM 1993; ISO 1987; OECD 1984; USEPA 1985). Issues that are specifically important for testing metals and metal compounds are discussed in greater detail. As indicated, aspects affecting both environmental and biological availability of metals are presented. B. Environmental Availability and Physicochemical Factors Culture Techniques. To initiate a toxicity test, algal cultures in log growth phase are needed. In most of the standard protocols considered in this review, very little guidance is given on the exact culturing techniques required to establish the test organisms in the log growth phase. However, this phase is of crucial importance to ensure the optimal and, perhaps more important in a regulatory context, reproducible growth and physiological condition of the organisms. Important factors for optimal growth include culture type, growth medium and nutrient concentrations, light conditions, temperature, and pH. Although the effects of these factors on toxicity test results are discussed in detail, it should be emphasized that pretest culturing variations can influence test results for the same physiological reasons as discussed next. In comparison to the studies examining factors affecting toxicity test results in the actual assay, little information is available on the effect of pretest culturing conditions. However, variation in culturing techniques of the algae before toxicity testing might well be an important factor affecting the wide variability of algal toxicity test results. Copper, for example, was found to be more toxic toward green algae under phosphorus limitation than under nitrogen limitation conditions (Hall et al. 1989a,b). Although these limitations refer to the actual test conditions, the type of limitation in the preculture medium could also be reflected in the outcome of algal tests. Until now, this possible source of test variation has not been examined. Most guidelines propose batch culture methods, although continuous culture techniques (e.g., chemostat) have been described (Hall et al. 1989b; Peterson 1991). In chemostat cultures, the test medium is being renewed during the exposure period and determines, together with a known limiting nutrient, the growth rate of the culture (Lederman and Rhee 1982). Hall et al. (1989b) demonstrated for copper that the use of continuous culture techniques instead of batch culture techniques (no renewal) alters the results of the subsequent algal toxicity tests.

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In the continuous chemostat culture, algae were stressed by nutrient limitation; subsequent Cu exposure demonstrated an increased sensitivity of the algae. Cells originating from a nutrient-rich batch culture that were exposed to a single pulse of Cu were less sensitive, however. The observation that algae are substantially more sensitive to Cu when exposed in chemostat cultures may raise concerns about the applicability to the natural environment of toxicity test results obtained from simple batch culturing. Problems associated with growth in batch cultures have been discussed in the context of basic algal physiology (Klaine and Ward 1983) and in the context of algal toxicity testing (Nyholm and Ka¨llqvist 1989). Culture and Test Media. Algae are normally cultured and exposed to the test substance in a nutrient-enriched medium. A large number of culture and test media for both freshwater and marine algae have been described (Guilard and Ryther 1962; Harrison et al. 1980; Stein 1973; Ukeles 1976), some of them, such as the Aquil and Fraquil media, actually being designed to simulate natural waters (Morel et al. 1979; Price et al. 1991). Most media used in standard protocols are prepared by adding prescribed amounts of macro- and micronutrients to distilled, filtered, or deionized water. The test medium, possibly with minor adjustments, is used as the culture medium. However, in most publications little or no information is given on the type of medium that has been used to culture algae. Unlike the media prescribed for toxicity testing, culture media are less well defined. It is clear that the type of culture medium will be reflected in the general health of the algal culture and might affect the physiological state of the algae and the outcome of toxicity tests. The importance of this factor cannot be ascertained as no literature is available. The effect of test medium composition on algal sensitivity has been studied extensively. Stauber and Florence (1989), for example, examined the toxicity of Cu and Zn to the marine diatom Nitszchia closterium and that of Cu and Pb to the freshwater alga Chlorella pyrenoidosa in two different media. They found that the ErC50 (growth rate as endpoint) for the freshwater species exposed to Cu increased from 24 µg/L in EPA-AAP medium (without ethylenediaminetetraacetic acid, EDTA) to >200 µg/L in a high-nutrient medium (MBL). Similarly, the toxicity of copper and zinc toward N. closterium changed with a factor of 10 to 20, depending on the test medium used. It is well recognized that test medium composition is a major factor affecting toxicity (Eloranta 1982; Millington et al. 1988). A comparison of essential experimental variables in current standard test protocols (ISO 1987; OECD 1984; USEPA 1978) including the medium characteristics, is given in Nyholm and Ka¨llqvist (1989). These authors identified pH, chelator presence, and major cation composition as important factors influencing toxicity. pH. Although the initial pH of standard algal tests is given in the respective guidelines, varying from 7.5 (USEPA 1978) to 8.3 (ISO 1987), considerable pH

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changes can occur during the test performance (Nyholm and Ka¨llqvist 1989; USEPA 1978). This change is a consequence of the interaction of three factors: nutrient composition (bicarbonate availability), algal growth, and CO2 exchange rate (aeration or shaking). Microalgae use primarily dissolved CO2 as their carbon source. In standard batch assays, the biomass density may quickly reach a level where the demand for CO2 by the growing algal population exceeds the transfer rate of CO2 from the gas phase to the liquid phase. Consequently, CO2 is also derived from the biocarbonate present in the standard media, varying between 15 and 50 mg/L NaHCO3. This utilization of biocarbonate leads to the formation of OH− ions and thus to an increase in pH. As the biological uptake rate of dissolved CO2 is saturated at very low CO2 concentrations, exponential algal growth continues until the bicarbonate source is exhausted, or until the pH rises to levels that adversely affect the algae. As population growth is reduced, pH may return to its initial value or may, through alkalinity-changing reactions (e.g., uptake of inorganic N), attain a higher or lower value. A full description of the mechanism is given in Nyholm and Ka¨llqvist (1989). Based on a large number of reported algal studies performed in static circumstances with no aeration and with high algal densities, these authors concluded that the algal populations probably were growth limited and experienced high pHs. The effect of shaking (increased CO2 availability) on algal growth and pH change in a standard USEPA S. capricornutum assay was described by Miller et al. (1978). They found that in the static test pH had increased from 7.5 to 9.5 after 3 d whereas in the shaken test the final pH had increased to 8.2. Population growth was clearly limited in the static test, exhibiting a linear growth after only 2 d. Considering the importance of even small pH changes on the speciation and toxicity of many metals, the observed differences caused by these simple operational variables may, and probably do, contribute to the variability of the standard algal toxicity results obtained for metals. Indeed, Peterson et al. (1984) reported that the toxicity of metals can change by a factor of 10 as a function of increasing pH. For Scenedesmus quadricauda, Cu toxicity even increased almost 76 fold (the concentration required to give 50% inhibition of phosphate uptake decreased from 7.6 to 0.1 µM/L) with an increase in pH from 5.0 to 6.5. pH effects on metal toxicity toward algae seem to be somewhat contradictory. Rai et al. (1981) reported an increased toxicity of Cu with lowered pH, associated with increasing Cu2+ activity. The opposite phenomenon, however, has been reported by many authors. Steeman-Nielsen and Kamp-Nielsen (1970) and Hargreaves and Whitton (1976) reported an increase of Cu toxicity with increasing pH for the algae Chlorella pyrenoidosa and Hormidium rivulare, respectively. There is growing evidence that this might be explained by the fact that hydrogen ions competitively exclude metal ions from binding to cell-surface ligands and hence inhibit metal uptake (Campbell and Stokes 1985; Parent and Campbell 1994; Peterson et al. 1984). Recently, Heijerick et al. (2002) and De Schamphelaere et al. (2003) found in their work with Zn and Cu that proton competition alone could not explain the observed increase of metal toxicity to

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P. subcapitata with increasing pH and suggested the presence of multiple binding sites that become available at elevated pH levels. This physiological effect of pH was reported earlier by Crist et al. (1988). Franklin et al. (2000a) examined the effect of pH on the toxicity of copper and uranium to the freshwater alga Chlorella sp. For copper they reported a decrease of the 72-hr EC50 from 35 to 1.5 µg/L when pH increased from 5.7 to 6.5. For uranium, 72-hr EC50s of 78 and 44 µg/L were found at pH 5.7 and 6.5, respectively. Macfie et al. (1994) also found a reduction of copper toxicity for Chlamydomonas reinhardtii (factor of 7) when pH decreased from 7 to 5. From the foregoing examples it is clear that pH control and carbon supply are, even in the current standard guidelines, factors that should be adequately monitored or controlled during the test. The use of most buffers (such as Tris) other than the presently used CO2/HCO3- buffer to control pH might not be appropriate because precipitation and complexation of the metals might occur. An exception might be the use of MOPS (3-[N-morpholino]propanesulfonic acid), which is reported to be completely noncomplexing for metals (Kandegedara and Rorabacher 1999) and which is recommended by USEPA (1991) because it does not change the toxicity of nontoxic effluents or the toxicity of toxic effluents and sediment pore waters. De Schamphelaere et al. (personal communication) also demonstrated that addition of MOPS up to a concentration of 1 g/L did not alter Zn or Cu toxicity to S. capricornutum in ISO medium (1987) within the useful pH range of this buffer whereas pH remained constant (±0.1 unit). Consequently, additional guidance in the present standard procedures is required to minimize pH as a source of variability. As low initial algal biomass and exposure periods of 2–3 d are already well defined and adequate in most standard procedures, strictly defining adequate mass transfer conditions for CO2, by using continuous shaking or aeration, might be a considerable step forward in controlling pH changes in routine algal assays. In conclusion, it is clear that metal toxicity toward algae cannot be separated from the pH of the test medium. The pH of standard test media is ±8 (EEC 1987; OECD 1984) and will therefore lead to low effects concentrations that are not representative for numerous natural surface waters with pH values lower than 7. From the foregoing, it also is apparent that an increase in pH (>±1) during the test period should be avoided when evaluating metal toxicity. To obtain a constant pH, several methods could be applied such as frequent adjustments of the pH with a nontoxic acid, through the use of a noncomplexing and nontoxic buffer such as MOPS (De Schamphelaere et al., (personal communication); Kandegedara and Rorabacher 1999; USEPA 1991) or by using airtight test flasks with CO2-enriched headspace (Halling-Sorensen et al. 1996). Although pH can be controlled with these techniques, one still might question the ecological relevance of CO2-limited test conditions. Chelators. The composition of the nutrient medium has been shown to affect the toxicity of chemicals such as pentachlorophenol and brominated organic

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compounds (Smith et al. 1987; Walsh et al. 1987b). Vasseur and Pandard (1988), however, reported that the medium composition did not affect metal toxicity toward S. capricornutum. The effect of one constituent, EDTA, has been investigated more than others. The EDTA concentrations present in OECD (1984), ISO (1987), USEPA (1985), and ASTM (1993) media are 0.1, 0.1, 0.3, and 0.3 mg/L, respectively. Low concentrations of EDTA are needed to complex trace metal micronutrients (e.g., Fe) in the culture medium. However, EDTA will also complex and reduce the toxicity of metal test substances if used in these media. If chelators are omitted, algal growth may be (nutrient) limited significantly during the test and log growth may be delayed or not attained (Lewis 1995). EDTA is not recommended for use in the test medium described by the USEPA methods for effluents (Weber et al. 1988, 1989), chemicals (USEPA 1985), and pesticides (Holst and Ellwanger 1982). Greene et al. (1991) reported on the effect of EDTA on the sensitivity of S. capricornutum to zinc. The effect of EDTA on the toxicity reduction of copper and cadmium was examined by Blaise et al. (1986) and Wren and McCaroll (1990). In their investigation into species-dependent variation of copper toxicity, Blanck et al. (1984) demonstrated, using a nonstandard test design, that copper sulfate toxicity in tests without EDTA ranged from 0.13 to 1.0 mg/L (14-d EC100s). Addition of EDTA to the medium (quantity not reported) changed the sensitivity range to 13–200 mg/L. It is clear that the presence of strong chelators such as EDTA will affect metal bioavailability and toxicity and should be taken into account when comparing metal toxicity data obtained in different test media. Nutrients and Other Medium Characteristics. In a batch culture, providing light absorption or CO2 mass transfer is not limiting, exponential growth will prevail at a constant rate until the limiting nutrient has been taken up. Growth may still continue until the intracellular reserves have been used up (Nyholm and Ka¨llqvist 1989). In comparison with phosphorus limitation, algal growth is more rapidly suppressed under nitrogen-limiting conditions. An in-depth study of the influence of the mineral composition of the medium on growth of planktonic algae was reported by Chu (1942). Comparisons of the nutrient composition of the standard algal culture and test media are given by Millington et al. (1988), Nyholm and Ka¨llqvist (1989), and Ho¨rnstro¨m (1990). According to these authors, although differences in composition do occur, in general these media are rather similar in both the type of nutrients and their quantity. Nitrogen concentrations range from 4 to <10 mg/L, whereas P concentrations between 0.19 and <0.7 mg/L are suggested. Concentrations of other medium characteristics such as bicarbonate, calcium, and magnesium can vary, according to the test protocols used, from 15 to 50 mg/L NaHCO3, from 1.2 to 4.9 mg/L Ca2+, and from ±0.7 to 1.5 mg/L Mg2+. Hardness of the algal medium is lowest (15 mg/L CaCO3) in the USEPA protocol and possibly highest in the OECD guideline (>60 mg/L CaCO3). Despite the apparent similarities among standard media, Adams and Dobbs (1984) found that medium composition had a significant effect on the toxicity of aminotriazole to S. capricornutum. Similarly, Mil-

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lington et al. (1988) examined the influence of three media (Bolds basal, OECD, and EPA) on the toxicity of morpholine, nitrilotriacetic acid, dichlorobenzene, and triphenyl phosphate to three algal species. In the tests with S. capricornutum, the 3-d LOEC values obtained in the different test media varied up to a factor of 10 (nitrilotriacetic acid). For Scenedesmus subspicatus, an identical maximum difference was observed, whereas for Chlorella vulgaris medium composition influenced toxicity values by a factor of 16. Metal toxicity to marine and freshwater algae may be regulated to some extent by macronutrient availability (Bates et al. 1985; Davies 1978). There are several reports of decreased toxicity in response to increases in P concentrations (Harding and Whitton 1977; Meijer 1972; Monahan 1973; Li 1979; Say and Whitton 1977). A possible explanation for these observations, based on the literature, is given by Hall et al. (1989a). High P concentrations in the medium result in elevated cellular P concentrations, in excess of immediate cell requirements, and are stored in polyphosphate bodies (Aitchison and Butt 1973; Rhee 1972, 1973). Polyphosphate bodies have been reported to act as a site for intracellular detoxification of metals (Pettersson et al. 1985). Enhanced cellular P levels may consequently lead to increased algal detoxification capacity and thus lower observed metal toxicity; this was also suggested by the findings of Twiss and Nalewajko (1992), indicating that internal polyphosphate plays a passive role in protecting Scenedesmus acutus from Cu. The role of nutrient limitation on Cu toxicity to algae was investigated by Hall et al. (1989a) in batch tests and in chemostat cultures (Hall et al. 1989b). In the first study, these authors observed that both Chlamydomonas geitleri and Chlorella vulgaris were more sensitive to Cu effects in batch cultures under P limitation than under N limitation. The growth rate of Chlamydomonas geitleri decreased to approximately 80% of that of the control in both the N- and Plimited cultures at an intracellular Cu concentration of 0.008, expressed as a Cu: C ratio. At a Cu:C ratio of 0.02 the growth rate under P limitation was reduced to 46% of the control whereas a reduction to 77% was observed in the N-limited cultures. Expressed in external Cu2+ activity, it was noted that in P-limited cultures of Chlorella vulgaris the division rate decreased steeply from 5.0 to 2.24 divisions d-1 between pCu 5.26 and 5.0. Similarly, in the N-limited cultures, a decrease from 2.4 to 1.62 divisions d-1 between pCu 5.07 and 4.9 was noted. For Chlamydomonas geitleri, growth rates decreased from 2.07 to 1.21 divisions d-1 at pCu below 5.2 in P-limited cultures and from 2.29 to 1.66 divisions d-1 in N-limited cultures. As well as explaining some of the variability in the copper algal toxicity reported in literature, the observed sensitivity differences in nutrient-limited algae might also have ecological implications. Quantities of micronutrients in test and culture media used in standard guidelines can vary considerably. For example, Zn concentrations in OECD (1994), ISO (1987), USEPA (1978), and ASTM (1993) media are 1.44, 1.44, 1.57, and 0.78 µg/L, respectively. Copper concentrations are 3.73, 3.73, 4.47, and 2.24 ng/L, respectively. The influence of these trace elements on the sensitivity of algae to environmental contaminants has not been examined in depth. This fac-

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tor may be especially important in toxicity assessments of metal and metal compounds, as it has been reported that preexposure to metals during culturing can affect the sensitivity of algae in (metal) toxicity tests. Muyssen and Janssen (2001) acclimated S. capricornutum and Chlorella vulgaris during 100 d to ISO medium containing 65 µg/L Zn and noted for both species an increase in Zn tolerance by a factor of 3 (72-hr EbC50 increased from 39 to 117 µg/L total Zn for S. capricornutum and from 34 to 105 µg/L total Zn for C. vulgaris). Compared to nonacclimated algae, acclimated S. capricornutum exhibited higher growth rates in all Zn treatments and in the controls, suggesting that the use of ISO medium results in suboptimal growth due to Zn deficiency. Indeed, because of the presence of EDTA, a free Zn2+ concentration of only 0.051 µg/L is calculated in the normal ISO-medium. An increased growth rate was not observed for C. vulgaris. Other authors (Coleman et al. 1971; Erre´calde et al. 1998; Jensen and Rystad 1974; Maeda et al. 1990) also found increased growth in the presence of moderate Zn concentrations. Erre´calde et al. (1998) suggested that a slight Zn deficiency in their test medium (Fraquil medium; Morel et al. 1975) could be at the base of these observations. There are other observations of adaptation to metal levels in culture. Stokes and Dreier (1981) demonstrated a loss of Cu tolerance in a Cu-tolerant isolate of Scenedesmus sp. after culturing the algae for 10 generations in Cu-free medium. Kuwabara and Leland (1986) found that Cu toxicity to S. capricornutum was reduced after 20 generations of Cu exposure. Finally, Ahluwalia and Kaur (1988) found that 0.1 and 0.5 mg/L cobalt chloride supported growth of C. vulgaris and the blue-green alga Anabaena variabilis. In conclusion, there are clear indications that nutrients and trace elements affect the sensitivity of algae to metals. For trace elements, the concentrations of these elements in the culture medium should be considered when evaluating toxicity data, especially those results that are used for water quality criteria derivations and risk assessments. Other Factors. Temperature and light need to be controlled during algal toxicity tests because they can significantly affect algal growth (Gaur and Singh 1991). In standard test guidelines, constant illumination is recommended at intensities varying from 30–90 (ASTM 1990) to 300 µE m-2s-1 (USEPA 1985) depending on the test protocol. Recommended temperatures are between 20° and 25°C and are species- and method specific (Lewis 1995). Current standard algal tests have exposure periods ranging from 3 to 4 d. Longer exposure periods using batch procedures should be avoided as algal growth may become nutrient limited or changes in the bioavailability of the contaminant (e.g., adsorption, degradation) may occur. Few data are available on the effect of these variables on metals toxicity toward algae. Only Mayer et al. (1998) found that the sensitivity of S. capricornutum to potassium dichromate was less under light limitation than under light saturation. Temperature only influenced the results indirectly by interacting with light (increase of the saturation intensity at higher temperature).

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C. Biological Availability and Biological Factors Test Species. Nalewajko and Olaveson (1998) reported that selection of the test species is the most important step in the performance of algal toxicity assays. Based on the arguments presented here, the present authors tend to agree with this statement, although species selection should depend on the purpose of the toxicity test. The algal species recommended in various standard methods are given in Table 2 . Many of these were chosen because of their availability and ease of culture and, to a lesser extent, because of their ecological relevance and sensitivity to toxicants (Lewis 1995). The species most frequently used is the green alga Selenastrum capricornutum Printz (renamed Pseudokirchneriella subcapitata; Hindak 1990). Other less commonly used species include the green algae Chlorella vulgaris, Scenedesmus quadricauda, and Scenedesmus subspicatus, and some diatoms, cyanobacteria, and chrysophytes. The predominance of Selenastrum as the standard test species can be attributed in part to its use in the AAP bottle tests described in the Introduction (Leischman et al. 1979). The infrequent use of diatoms, cyanobacteria, and chrysophytes in toxicity tests is due to the difficulties encountered in maintaining cultures, lack of commercial sources, and difficulty in obtaining sufficient growth during the 3- to 4-d test period. Increasing concern has been expressed, by researchers and regulators alike, about the reliance on Selanastrum capricornutum as the “almost exclusive” test species representative of all microalgal species and communities (Blanck et al. 1984; Klaine and Lewis 1994; Lewis 1990; Nyholm and Kallqvist 1989; Wangberg and Blanck 1988). From the point of view of ecological relevance, important for certain uses of algal toxicity data, S. capricornutum is not as widely

Table 2. Overview of algal species used in common standard algal methods. Species Chlorophyta Selenastrum capricornutum Printz Scenedesmus subspicatus Chodat Scenedesmus quadricauda (Turp.) Breb. Chlorella vulgaris Beij. Cyanophyta Microcystis aeruginosa Kutzing Anabaena flos-aquau (Lyng.) Breb. Chrysophyta Cyclotella sp. Navicula pelliculosa Grunow Nitzschia sp. Synedra sp.

Method

ASTM, APHA, OECD, ISO, USEPA OECD, ASTM, ISO USEPA OECD, ASTM, USEPA ASTM, APHA ASTM, APHA APHA ASTM, APHA APHA APHA

References: APHA 1985; ASTM 1993; ISO 1987; OECD 1984; USEPA 1985.

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distributed in the environment as perceived. For example, Riemer (1984) reported that although the genus is widely distributed in U.S. waters, S. capricornutum is not indigenous to this country. Greeson (1982) noted that, of 500 genera of freshwater algae present in the U.S., 58 were dominant. Selenastrum was not considered dominant but is classified as commonly occurring. More important, reliance on one species for risk assessment may over- or underpredict toxicity to other species that exhibit different toxicokinetics. The majority of freshwater microalgae used in toxicity testing belong to one of eight possible classes. As this classification is made on the external and internal characteristics of the algal cell, it should be clear that all these classes, genera, or species might very well have different responses to contaminant uptake and ensuing toxicity (Bold and Wynne 1985). This variation may be particularly important in species for which exclusion of metals is the only operative mechanism of metal resistance, as demonstrated by Hawkins and Griffiths (1982) for four species of marine phytoplankton exposed to Cu. Nalewajko and Olaveson (1998) further reviewed the ecology of natural algal assemblages in the context of species representativeness of standard algal toxicity tests. It should be emphasized that no validation studies have been performed that examine the potential of any algal species as surrogate test species in laboratory or field studies. Perhaps the most important argument for increasing the number of species routinely used in algal assays is provided by the extensively reported intra- and interspecies differences in sensitivity of algae toward environmental contaminants (Blanck et al. 1984; Bringmann and Ku¨hn 1978; Fizgerald and Faust 1963; Fitzgerald 1964; Lewis and Hamm 1986; Yamane et al. 1984). Lewis (1995) noted that “the importance of using other species (than S. capricornutum) cannot be overstated, because the interspecific variation in the response of algae to chemicals can vary over several orders of magnitude.” A consistently sensitive algal species has not been identified (Lewis 1995). From the current literature it is unclear whether the reported differences in sensitivity among various taxa, and indeed within individual species, are real or artifacts caused by differences in culturing conditions, media composition, or other test abiotic variables. Next to these factors, the extent to which the physiological state of the algae during and before the assay affects the results is rarely considered or studied. It is hoped that this review can contribute to the identification of these “neglected” biological concepts. Examples of interspecific differences in algal sensitivity are summarized in Lewis (1995). In toxicity tests with copper sulfate, Fitzgerald and Faust (1963) compared sensitivity of Microcystis aeruginosa and Chlorella pyrenoidosa and observed a difference of a factor of 20. Similarly, USEPA (1980) reported a toxicity difference between 80 and 150 when comparing results from Cu assays with the frequently used regulatory species Chlorella pyrenoidosa, Scenedesmus quadricauda, Selenastrum capricornutum, and Chlamydomonas sp. Bringmann and Ku¨hn (1978) noted a sensitivity difference of 300, 300, and 37 for Microcystis aeruginosa and Scenedesmus exposed to sodium dichromate, nickel chloride, and copper sulfate. Using four green and blue-green species, Vocke et al.

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(1980) found differences of 640, 4, 7, and 257 for As, Cd, Hg, and Se. Based on the Zn toxicity data reported by Les and Walker (1984) and Van Ginneken (1994), an eightfold difference in sensitivity was calculated between S. capricornutum (3-d NOEC of 24 µg Zn/L) and Chroococcus paris (10-d NOEC of 200 µg Zn/L). For cadmium, 3-d NOECs are reported ranging from 8.5 µg/L for Fragilaria crotonensis (Conway and Williams 1979) to 31 µg/L for Scenedesmus subspicatus (Ku¨hn and Pattard 1990). Wong et al. (1979), on the other hand, reported an 6-d EbC50 of 1000 µg Cd/L for Ankistrodesmus falcatus. Blanck et al. (1984), using the same microplate test design and test media to avoid sensitivity differences caused by the test setup, tested 19 chemical on 13 freshwater algal species. They demonstrated that a species-dependent variation in algal sensitivity, based on 14-d EC100 values, may exceed three orders of magnitude depending on the chemical tested. The EC100s observed for copper sulfate ranged from 0.13 to 1.0 mg/L, for mercuric chloride from 0.063 to 0.5 mg/L, and for potassium dichromate from 0.13 to 16 mg/L. The authors further concluded that no generally sensitive algal species could be identified. However, based on the distribution of sensitivity toward the 19 chemicals, some algae seemed to be more extreme in their response to chemicals. Scenedesmus obtusiusculus, Chlorella emersonii, and Selenastrum capricornutum were classified as “more representative” in their sensitivity because of their overall narrow sensitivity distribution. Monodus subterraneus, Raphidonema longistra, and Synechococcus leopoliensis, on the other hand, were very sensitive to some and very insensitive to other compounds. These observations are clearly in contradiction with the conclusions of Roberts et al. (1982), who reported that differences in algal sensitivity were small and who consequently concluded that a single standard species could be representative of an entire taxon. It should be noted that the latter authors only tested three algal species and three chemicals. In general, the observation of large differences in interspecies sensitivity supports the suggestion of various authors (Blanck et al. 1984; Wangberg and Blanck 1988) to use a battery of algal species from a broad range of taxonomic representation in the laboratory testing process, thus increasing ecological realism. Which algal species to be used in a test battery has not been suggested, but it is obvious that selected algal species should be representative for the area of concern. Tables 3 and 4 summarize reported Zn and Cd toxicity data, respectively, for different species and different endpoints. Next to interspecies variation, intraspecies variability in standard algal toxicity test results also is considerable. For instance, Starodub et al. (1987) reported 2-wk NOEC for growth ranging from 100 to 230 µg Zn/L for Scenedesmus quadricauda. Tables 3 and 4 also present reported differences in Zn and Cd toxicity for Selenastrum capricornutum. For Zn, a factor of 6 is observed between the lowest and highest effect concentration (endpoint, growth rate EC50). For Cd a factor of 100 was found for the same species and the same endpoint. As with the observed variability between species, it unclear if these differences are experimental artifacts or real differences between clones. Interlaboratory round-robin testing of regulatory algal tests has demonstrated that a large

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37

Table 3. Reported literature toxicity data for zinc.

Species Selenastrum capricornutum

Endpoint 3-d NOEC (g)

4-d NOEC (g) 2-wk NOEC Growth rate, EC50

Scenedesmus quadricauda Chlorella vulgaris Croococcus paris

2-wk NOEC (g) 4-d NOEC (g) 2-wk NOEC (g) 5-wk NOEC (g) 10-d NOEC (g)

Effect concentration (µg Zn/L)

Reference

50 24 8 10 5 63 10 67 100 230 1200 400 560 200

Van Woensel 1994 Van Ginneken 1994 LISEC 1997 Bartlett and Rabe䉳 1974 Kuwabara 1985 Wolterbeek et al. 1995 Kuwabara 1985 Erre´calde et al. 1998 Starodub et al. 1987 Starodub et al. 1987 Bringmann and Ku¨hn 1978 Ahluwalia and Kaur 1988 Rosko and Rachlin 1977 Les and Walker 1984

variability of toxicity test results performed within the same species is not uncommon (Hanstveit 1982; Ka¨llqvist et al. 1980). Nyholm (1985) states that EC50 estimates, obtained with identical species, could differ between laboratories by a factor of more than 1000. The lack of comparability between different methods seems primarily to be due to a number of physicochemical factors that either differed between methods or which were inadequately controlled (Ka¨llqvist 1982; Nyholm 1982). Effects Measurements. The effects criterion most frequently reported is growth inhibition based on changes in biomass EbC50 as a function of toxicant concentration. Another response variable prescribed by some standard guidelines is the growth rate (ErC50) (ASTM 1993; OECD 1984). The indiscriminate use of algal toxicity data based on both types of endpoints for regulatory purposes has led to an unresolved scientific controversy (Lewis 1995; Nyholm 1985). Nyholm (1985) made an in-depth study of the technical and mathematical aspects of this issue and concluded that, from a theoretical point of view, growth rate is a better response variable than biomass, e.g., EC figures based on growth rates are less dependent on particular test system parameters. Realizing that both types of data analysis probably will continue to be used, Nyholm underlined the importance of making scientists and regulators aware of the potential differences between both types of EC estimates. In current regulatory practice, however, the concerns and recommendations expressed by Nyholm and others have been largely ignored and, in most cases, algal data continue to be used as reported in databases, i.e., with little consideration for the endpoint reported.

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C.R. Janssen and D.G. Heijerick

Table 4. Reported literature toxicity data for cadmium. Species

Endpoint

Selenastrum capricornutum

Cell number, EC50 Growth rate, EC50

Biomass, NOEC Biomass, EC50

Scenedesmus quadricauda Scenedesmus subspicatus

Average cell number, EC50 Biomass, NOEC Biomass, EC50

Growth rate, EC50 Chlamydomonas Steady-state cell reinhardtii number, NOEC Chlorella vulgaris Cell number, EC50 Biomass EC50 Chlorophyll a content, EC50

Effect concentration (µg Cd/L)

Reference

23 18 89 79 70 120 13 341 32 5.6 3.4 6.7 62 50 57 45 6.1

LISEC 1998a LISEC 1998b LISEC 1998a LISEC 1998b Janssen 1993a Janssen 1993b Chen and Lin 1997 Chen and Lin 1997 Lin et al. 1996 Sedlacek et al. 1983 Chiaudani and Vighi 1978 Chiaudani and Vighi 1978 Erre´calde et al. 1998 Bartlett and Rabe 1974 Turbak et al. 1986 Erre´calde et al. 1998 Klass et al. 1974

31 62

Bringmann and Ku¨hn 1978 Ku¨hn and Pattard 1990

136 7.5 60 1220 550 393

Ku¨hn and Pattard 1990 Lawrence et al. 1989 Rosko and Rachlin 1977 Jouany et al. 1983 Jouany et al. 1983 Kosakowska et al. 1988

The relative merits of these and other effects parameters and techniques to measure biomass (e.g., cell counts, dry weight, chlorophyll a measurements) are also discussed in Nyholm (1985), Nyholm and Ka¨llqvist (1989), and Nalewajko and Olaveson (1998). Inoculum. The number of algal cells at the start of the test can affect the outcome of toxicity assays with metals and contaminants in general. High initial cell densities will influence the medium through increased photosynthesis, increased numbers of algal cells, and increased concentration of degradation products of these cells. An inoculum of a magnitude of approximately 5 × 102 cells/

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mL in a 72–96 hr toxicity test will minimize the aforementioned effects so the algal cells can develop in a medium that remains more or less the same during the test period, including the metal concentration (Ho¨rnstro¨m 1990). This proposed cell density is two orders of magnitude lower than the inoculum cell densities required in standard guidelines (ASTM 1990; ISO 1987; OECD 1984; USEPA 1985). The importance of cell density has also been pointed out by Franklin et al. (2000b), who concluded that Cu toxicity to Chlorella sp. decreased with increasing initial cell concentrations (1.102 –1.105 cells/mL). They demonstrated that a decrease of the equilibrium concentration of dissolved Cu in test medium with high initial cell densities was due to binding to cell surfaces and not the result of increased algal exudate production. Their findings suggest that in standard algal bioassays, using 1.104 –1.105 cells/mL, the algae themselves alter the partitioning and the speciation and therefore these bioassays may underestimate metal toxicity. Regardless of these conflicting reports, in general the initial algal population should grow optimally and exponentially (Ho¨rnstro¨m 1990). Data Analysis. Walsh et al. (1987a) evaluated and compared four different statistical methods for deriving algal EC50s (straight-line graphic interpolation, moving average interpolation, probit analysis, and the binomial method) and found that resulting EC50 values were essentially identical. In a discussion of the problems of data treatment, Adams et al. (1985) suggested a multiple comparison method (based on Dunnett’s test) to calculate the LOECs. It is clear, however, that the reliability of the calculated effect concentration will be dependent on the observed dose–response curve: significant differences can be found between the outcome of different statistical methods when the 50% effect is situated at the borders of the tested concentration range.

IV. Adsorption, Uptake, and Toxicity Mechanisms for Metals There is a large discrepancy between the large body of fundamental knowledge on toxicity and toxicity mechanisms of chemicals in general, and metals in particular, and the manner in which routine algal toxicity data are aquired and used in regulatory contexts. The brief discussion given next presents some of the fundamental principles that are well established but have not yet been considered in standard algal testing of metals. Many qualitative data are available demonstrating that the total aqueous concentration of a metal is not a good predictor of its bioavailability (Campbell 1995). As pointed out earlier, differences in bioavailability of metals during standard algal testing can be an important factor in determining the large variability in algal test results with metals. The fact that this issue confounds correct application of algal toxicity data in certain regulatory arenas is surprising because the free ion activity model (FIAM), explaining metal–organism interactions, has been available since the early 1980s (Morel et al. 1979). Many of the

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early accumulation and toxicity experiments contributing to the development of this model were carried out with unicellular algae. Two examples are now used to illustrate the concept and importance of this model. Anderson et al. (1978) examined the importance of Zn as an essential metal for the growth of the coastal diatom Thalassiosira weisflogii. When the free Zn activity was manipulated by increasing the total Zn concentration at each of three different EDTA concentrations, growth increased as Zn was added, thus demonstrating that the algae were indeed Zn limited. Replotting the same growth data as a function of the free Zn ion activity, all three EDTA treatments showed the same dependence of the growth rate on the [Zn2+], clearly demonstrating the Zn ion activity rather than the total Zn concentration determined the growth rate under Zn-limiting conditions. Examining the toxicity effects of Zn on algal growth, Allen et al. (1980) added different ligands to the AAP test medium of the blue-green alga Microcystis aeruginosa. They observed a reduction in Zn toxicity in cultures with added chelators; without ligands, algal growth was completely inhibited. Plotting the log yield as a function of the calculated free Zn ion concentration resulted in a linear relationship, even though the conditional stability constants of the five chelators varied by nine orders of magnitude and their concentrations by three orders of magnitude. Again, the effects of Zn on the algae were directly dependent on the concentration of the free metal ion rather than the total metal concentration. Bates et al. (1982) demonstrated the same phenomenon for the adsorption and uptake of Zn by Chlamydomonas variabilis and Scenedesmus subspicatus. Similar uptake, adsorption, and toxicity experiments with Cu, Zn, Ni, Mn, and Fe have been reported using the following freshwater species: Chlorella vulgaris, Oocystis marssonii, Scenedesmus quadricauda, Ankistrodesmus falcatus, Scenedesmus subspicatus, Chlorella variabilis, and Chlamydomonas reinhardii (Hall et al. 1989a; Manahan and Smith 1973; Petersen 1982; Peterson and Healey 1985; Spencer and Nichols 1983; Schenck et al. 1988; Xue et al. 1988). From his extensive review on the use of FIAM to explain metal–organism interactions, Campbell (1995) concluded the following. Of the 59 studies performed at constant pH and hardness and in the absence of natural dissolved organic matter (DOM), the biological response in 52 of those varied as a function of the free metal ion as predicted by FIAM. Concerning the studies performed in the presence of natural DOM, Campbell (1995) noted that these studies were contradictory: there were very few quantitative studies available to evaluate the applicability of FIAM, and half of them supported the model while the other half seemed to be in contradiction. Campbell thus concluded that the applicability of the FIAM in natural waters remained to be demonstrated. It is surprising that despite our good fundamental knowledge of the effects of metals in laboratory toxicity, little has been done to incorporate this information into standard tests used for regulatory use. Indeed, no recommendations on how to keep the pH constant during the test period, using noncomplexing buffers, or the possible effects of present chelators (e.g., EDTA) on metal toxicity

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have been incorporated in the main standard guidelines (ASTM 1993; ISO 1987; OECD 1984). Changes of 1 pH unit, considered to be acceptable in most standard guidelines, may change the free metal ion fraction significantly, thus altering metal toxicity during the test period. For example, based on speciation calculations with the WHAM equilibrium model (Tipping 1994), Cu2+ concentrations in a standard medium are found to decrease by more than one order of magnitude when the pH increases from 7 to 8. Although modeling free metal activity in culture media may be more of a challenge than in simple ionic solutions or those where EDTA complexes dominate (Jackson and Morgan 1978), it is possible to compute the free metal activity from chemical models. In some culture media, however, where a variety of metal adsorbents such hydrated oxides of iron, manganese, and silicon are present, such calculations could be too inaccurate to be of use (Stauber and Florence 1989).

V. Acclimation and Adaptation Under Laboratory Conditions Individual organisms can increase their tolerance to stress by acclimation, adjusting physiological or biochemical mechanisms. Populations can evolve stress tolerance through gene selection (adaptation). An example of acclimation is the production of phytochelatins, metal-binding peptides, as a response to metal stress (Gekeler et al. 1988; Grill et al. 1985; Howe and Merchant 1992). These phytochelatins are considered to play an important role in metal homeostasis and detoxification (Grill et al. 1985; Rauser 1990). Although it is sometimes difficult to distinguish between the two phenomena, there is an increasing body of evidence for rapid genetic evolution resulting from a variety of anthropogenic stresses. Metal tolerance in algae has been described by Butler et al. (1980), Chapman (1985), Fitzgerald and Faust (1963), Klerks and Weis (1987), Silverberg et al. (1976), Stockner and Antia (1976), and Stokes and Dreier (1981). Soldo and Behra (2000) exposed a natural periphyton population from the river Glatt (Switzerland) during 16 wk to various Cu concentrations (ranging from 0.05 to 5 µM) and observed an increased Cu tolerance in populations previously exposed to Cu. Additionally, these communities also displayed a cotolerance to Zn, Ni, and Ag. Acclimation or adaptation of laboratory test organisms to deficient, suboptimal, or elevated trace metal concentrations, especially those species with short generation times such as algae, might have important, unrecognized environmental management implications. Laboratory studies have already demonstrated the possibility of algae acclimating to elevated metal concentrations (Muyssen and Janssen 2001), but the observed increased tolerance to Zn disappeared rapidly when the algae were returned to their original culture conditions. Evidence of genetic adaptation of algae is given by Klerks and Weis (1987), who compared metal toxicity data from algal populations originating from different polluted sites. They concluded that the observed increases in tolerance toward metals has in most cases a genetic component. In general, resistance differences in

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bacteria, algae, and fungi were shown to have a genetic basis. For example, Jensen and Rystad (1974) demonstrated genetically determined differences in Zn tolerance between two clones of the alga Skeletonema costatum, the more tolerant being obtained in a more heavily polluted fjord. Similar observations were made by Say et al. (1977) for Zn in the alga Hormidium rivulare and by Foster (1977) who examined Cu tolerance in the alga Chlorella vulgaris. The acclimation/adaptation phenomenon is probably underestimated in laboratory studies, given the diversity in procedures, test species, and media that are currently used. The origin of the species, or the time that this species has been cultured in a specific medium, may lead to significant differences in observed effect concentrations obtained with the same species in different laboratories, thus adding to reported intraspecies sensitivity differences.

VI. Laboratory to Field Extrapolation When extrapolating laboratory data to field circumstances, one must realize that natural waters are chemically very complex. For example, DOC present in natural waters is known to reduce metal toxicity through complexation. The concentration of Ca2+ and Mg2+ ions in some natural waters also may be very different from that in standard media and may affect metal toxicity. No data are reported on the influence of elevated hardness on metal toxicity to algae. Smith et al. (1987) reported a decrease of the toxicity of pentachlorophenol (PCP) to Selenastrum capricornutum by a factor of 7 when hardness, as CaCO3, increased from 21 to 126 mg/L. Another parameter that must be considered is the nutritional state of natural field waters; possible deficiency or excess of one or more nutrients, both major nutrients and trace elements, may lead to a suboptimal physiological state of the natural phytoplankon community, which could affect their sensitivity to various contaminants. Until now, very few field toxicity studies have been performed with local algae. Considering the foregoing discussions of (sensitivity and methods, modifying factors, and intercalibration exercises, it is clear that the toxicity of metals to algal communities in the field will depend on the species present and on environmental factors that modify toxicity.

Summary Current regulatory methods for assessing the effects of contaminants, and metals in particular, rely mainly on a limited number of standardized test methods and test species (OECD, ISO, ASTM, USEPA). However, these test protocols allow a certain degree of freedom in relation to physicochemical parameters or biological aspects, which may lead to large variability in test results. The current review, based on effects data and theoretical considerations reported in the literature, tried to determine and quantify the effect of variation of these factors on the outcome of metal toxicity tests with algae. Major physicochemical parameters that affect metal toxicity to algae are hardness, pH, preculture conditions,

Table 5. Overview of the test variables potentially affecting metal toxicity in standard algal methods and recommendations for improving current standard test procedures, more specifically for metal toxicity testing.

Test variable

Effect

Preculture conditions

Changes in physiological state of algal cells

Nutrient limitation

Changes in sensitivity as a function of type of limitation Affecting metal toxicity through speciation changes, Me2+/H+ competition, and possible structural changes of the algal membrane

pH

Illumination Type of test species Adaptation/acclimation to higher or lower background concentrations

Changes in metal toxicity and bioavailability through complexation Decrease of sensitivity under light limitation Variation through interspecies sensitivity differences Intraspecies variability in sensitivity

High initial cell concentrations may lead to an underestimation of metal toxicity

Endpoint

Different endpoints (growth rate, biomass, etc.) will result in different effect concentrations

43

Inoculum

Clear definition of preculture conditions and culture test media Reduce growth rates and/or reduction of initial cell concentration at beginning of the test pH control (variation of 0.1–0.2 pH units) through: Reduced growth conditions and/or reduction of initial cell concentration at beginning of the test Addition of a nontoxic and noncomplexing buffer CO2-enriched headspace Replacement of EDTA by a less metal-complexing compound (e.g., artificial humic acid) Definition of a narrow range of allowed light intensity Testing with a limited number of (defined) algal species covering a broad sensitivity range Clear definition of culture test media, preculture conditions, and acclimation period of algal population to these conditions Reduction of initial cell concentration at beginning of the test (pH control, reducing nutrient limitation) Effects on both growth rate and biomass should be reported and the statistical treatment methodology should be defined

Algal Toxicity Tests

Chelators

How to improve current standard test procedures?

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type of test medium, and presence of chelating agents. Literature data also clearly demonstrate the importance of test species or strain selection (inter- and intraspecies sensitivity variability) on the outcome of algal toxicity tests. For Zn, a factor of 8.3 is observed between the NOEC for Selenastrum capricornutum (currently renamed Pseudokirchneriella subcapitata) and Croococcus paris. An intraspecies difference for S. capricornutum of a factor of 60 is observed between various reported EC50s for Cd. Next to differences in physicochemical test conditions, possible adaptation or acclimation to deficient/elevated metal concentrations add to the reported differences: S. capricornutum became three times less sensitive to Zn when acclimated to 65 µg Zn/L compared to cultures in ISO medium. This review has revealed that currently accepted standard protocols used in regulatory frameworks contain a number of major shortcomings on the physicochemical and biological aspects of algal toxicity testing with metals. These shortcomings are summarized in Table 5, together with a number of suggestions that could help to modify and improve standard test protocols for evaluating metal toxicity to algae. Until now, important factors such as pH control during test performance, selection of test medium, test species, and the effects of possible adaptation/acclimation to natural metal concentrations have not been considered, which could have serious implications when the resulting unsuitable or irrelevant toxicity data are subsequently used for setting environmental management policies. These findings also have their consequences when extrapolating laboratory data to the field as the complexity of natural waters currently is not reflected in laboratory standard media. These media contain no dissolved organic matter, have a relatively high pH, and contain large amounts of essential nutrients. In addition, the limited number of laboratory test species do not reflect natural phytoplankton communities. Test procedures for assessing the environmental impact of metal contamination in a specified ecoregion should therefore be based on performing a battery of algal tests with species adapted to and tested under the specific natural conditions of the region.

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 Springer-Verlag 2003

Rev Environ Contam Toxicol 178:53–91

Chemistry of Chromium in Soils with Emphasis on Tannery Waste Sites S. Avudainayagam, M. Megharaj, G. Owens, R.S. Kookana, D. Chittleborough, and R. Naidu Contents I. Introduction ....................................................................................................... II. General Chemistry of Cr .................................................................................. A. Position in Periodic Table ........................................................................... B. Stable Oxidation States ................................................................................ C. Chemistry of Cr(VI) .................................................................................... D. Chemical Reactivity of the Chromate Ion .................................................. E. Chemistry of Cr(III) ..................................................................................... III. Toxicity of Cr .................................................................................................. A. Toxicity to Plants and Microorganisms ...................................................... B. Toxicity to Animals and Humans ............................................................... C. Carcinogenicity and Genotoxicity ............................................................... IV. Chemistry of Soil Cr ........................................................................................ A. Chromium Cycle in Soil and Water ............................................................ V. Solid-Phase Speciation of Cr in Native Soils .................................................. A. Solid-Phase Speciation of Cr in Sludge and Sludge-Amended Soils ........ B. Solid-Phase Speciation of Cr in Industrial Contaminated Soils ................. C. Solid-Phase Speciation in High-Cr-Contaminated Soils ............................. D. Partitioning and Mobility of Cr ................................................................... VI. Factors Affecting the Soil Solution Chemistry of Cr ...................................... A. Soil Solution pH .......................................................................................... B. Chromium (III) in Soil Solution .................................................................. C. Solubility of Organically Complexed Cr(III) .............................................. D. Adsorption and Desorption of Cr(III) ......................................................... VII. Oxidation of Cr(III) to Cr(VI) ......................................................................... A. Effect of pH ................................................................................................. B. Solubility of Cr(III) and Oxidation .............................................................

54 55 55 55 57 57 57 58 58 58 59 59 61 61 62 64 65 66 67 67 67 69 70 71 71 71

Communicated by George G. Ware. M. Megharaj, G. Owens, R. Naidu ( ) (formerly CSIRO Land and Water, Adelaide), Australian Centre for Environmental Risk Assessment and Remediation, University of South Australia, Mawson Lakes, Mawson Lakes, SA 5095, Australia. R.S. Kookana CSIRO Land and Water, Waite Road, Urrbrae, Adelaide SA 5064, Australia. S. Avudainayagam, D. Chittlebrough Department of Soil and Water, The University of Adelaide, Waite Campus, Glen Osmond, SA 5064, Australia. S. Avudainayagam Tamil Nadu Agricultural University, Coimbatore, India 641003.

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VIII. Factors Affecting Soil Solution Cr(VI) ........................................................... A. Effect of pH ................................................................................................. B. Adsorption of Cr(VI) ................................................................................... IX. Contaminated Soil Environment and Mobility of Cr(VI) ............................... A. Tannery Sludge Composition ...................................................................... B. Soil Solution Composition and Mobility of Cr(VI) .................................... X. Remediation of Chromate-Contaminated Soil and Water ............................... A. Remediation Strategies for Contaminated Soil ........................................... B. Remediation of Cr(VI)-Contaminated Water .............................................. XI. Future Research Needs ..................................................................................... Summary .................................................................................................................... Acknowledgments ...................................................................................................... References ..................................................................................................................

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I. Introduction Chromium (Cr), the 17th most abundant element in the earth’s mantle, is used in many industrial processes such as plating, alloying, tanning of animal hides, water corrosion inhibition, textile dyes and mordants, pigments, ceramic glazes, refractory bricks and pressure-treated lumber. Chromium also the trace element responsible for the brilliant red and green colors of the ruby and emerald (Petrucci and Harwood 1993). Chromium has attained wide public and regulatory attention because of its toxicity to environmental ecosystems, including both microorganisms and animals, under certain oxidation states. Its oxidation states vary between −2 and +6, but only the +3 and +6 states are stable under commonly observed environmental conditions (Cotton and Wilkinson 1980). These two oxidation states have different toxicity and mobility. Hexavalent chromium [Cr(VI)] is a known human carcinogen, via inhalation, and is mobile, whereas trivalent chromium [Cr(III)] is comparatively less toxic and relatively immobile (Yassi and Nieboer 1988). Elevated soil and water concentrations of this potentially harmful element principally result from industrial waste or spills. Tanning of leather with Cr salts [Cr(III)], first introduced in 1858, consumes considerable quantities of Cr (Thorstensen 1976). In fact, some of the earliest processes for chrome tanning of leather used Cr(VI) chemicals to saturate the skin and then reduce them to insoluble forms. However, the current standard practice now is to use a soluble trivalent compound and a masking agent, a procedure that allows Cr to effectively penetrate the hide (Lollar 1980). During this process, Cr(III) forms cross-links between the collagen fibers and gives leather its durable finish. The wastewater discharged from tannery industries contains heavy loads of both organic and inorganic chemicals that are typically different from other Cr polluting industries (Reemtsma and Jekel 1997). In the past three decades, as a result of an increased awareness and protection of environmental health, strict environmental regulations have been enacted to prevent indiscriminate disposal of tannery waste into soil and water bodies in the developed and developing parts of the world. However, before these regulations were enacted

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many tanning industries had commonly dumped their Cr-containing solid and liquid wastes to neighboring land and water bodies. The exact number of such contaminated sites would be difficult to quantify, and no published reports are available listing details of such sites in the world. However, an overview of tanning industries and the extent of contamination is available for Southeast Asian countries (Naidu et al. 2000a). The development of analytical techniques to speciate aqueous Cr, and reports of groundwater contamination of Cr(VI) in the vicinity of old tannery industries, have increased the attention of scientists toward historical sites where tannery waste has been disposed. Although Cr(III) is the predominant species discharged in tannery solid and liquid wastes, the presence of Cr(VI) in groundwater near old tanning industries raised critical questions about the thermodynamic stability of Cr(III). However, in natural systems, manganese oxides can oxidize Cr(III) to Cr(VI) (Bartlett and James 1979; Nakayama et al. 1981; Amacher and Baker 1982; Eary and Rai 1987; Saleh et al. 1989; Johnson and Xyla 1991; Fendorf and Zasoski 1992). A number of highly contaminated tannery waste disposal sites exist in countries throughout the world. This chapter presents an overview of the fate and behavior of Cr under different soil and climatic conditions and the implications to tannery waste-contaminated sites.

II. General Chemistry of Cr A. Position in Periodic Table Chromium, with an atomic number of 24 and mass number of 51.9961, belongs to the first series of transition metals. The elements vanadium, manganese, and molybdenum surround its position in subgroup VIB of the periodic system. Its electronic configuration is (Ar) 3d54s1. B. Stable Oxidation States Chromium exists in a number of oxidation states of variable stability. It is clear from the reduction potential diagram constructed by Neiboer and Jusys (1988) (Fig. 1) that in solution Cr(III) is the most thermodynamically stable species, because in solution considerable energy would be required to convert it to either lower or higher oxidation states. Although CrO2− 4 is also relatively stable, its high positive reduction potential means that it is also a strong oxidizing agent and therefore unstable in acid solutions in the presence of electron donors such as Fe2+, H3AsO3, and HSO−3 or organic molecules with oxidizable groups (alkanes, alkenes, alcohols, aldehydes, ketones, carboxylic acids, mercaptans, etc.) (Wiberg 1965; Beattie and Haight 1972). As Cr(III) and Cr(VI) are the most stable species of Cr in the environment, the chemistry of these species is discussed in greater detail in the following sections.

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Fig. 1. Reduction potential diagram for chromium. (From Neiboer and Jusys 1988.)

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C. Chemistry of Cr(VI) Hexavalent Cr is a strong oxidizer and consequently Cr(VI) exists only as pHdependent soluble oxygenated species governed by the following equilibria (Nieboer and Jusys 1988): H2CrO4 ⇔ H+ + HCrO−4 − 4

+

HCrO ⇔ H + CrO

2− 4

log (Ka1) = 0.6 log (Ka2) = −5.9

Because in environmental matrices the pH varies only between 3 and 10, HCrO−4 and CrO2− 4 are the dominant species present in natural systems. In addition, at concentrations of Cr(VI) greater than 0.01 M, dimerization of the chromate ion occurs (Beas and Mesmer 1976), yielding the dichromate ion (Whitten and Gailey 1984): − Cr2O2− 7 + H2O ⇔ 2HCrO4

log (K) = −2.2

In biological systems, where the chromate concentrations are below 0.01 M the existence of dichromate is not expected to be significant, especially at physiological pH values of 7 to 8. D. Chemical Reactivity of the Chromate Ion Chromic acid has been used extensively as an oxidizing agent in both organic and inorganic chemistry, and comprehensive reviews on these subjects have been published elsewhere (Wiberg 1965; Espenson 1970; Beattie and Haight 1972). Discussing the oxidizing power of chromate, Nieboer and Jusys (1988) concluded that chromate is chemically unstable at physiological pH values, but does not decompose as easily as expected on the basis that it has a highly positive standard reduction potential and a biological half-life in hours. The chromate ion is kinetically nonlabile because it is difficult to displace the oxygen atoms from the chromate ion (CrO42−). In contrast, the hydroxyl group of the monoprotonated form of chromate (HCrO4−) is more readily replaced. E. Chemistry of Cr(III) Trivalent Cr species are generally considered to be nonlabile because ligand displacement is slow (hours to days at room temperature) compared to most other metal ions (10−9 –10−3 sec at room temperature) (Cotton and Wilkinson 1980). Due to this kinetic inertness, many Cr(III) complex species can be isolated that are stable in solution. As other trivalent metal ions, namely, Fe(III) and Al(III), the hydrated Cr(III) ion, Cr(OH2)3+ 6 , has a tendency to hydrolyze and this step is often accompanied by polymerization. Hydrolysis involves the conversion of a bound water molecule to the hydroxide ion and results in the release of a proton. Equilibrium measurements have identified the existence of 5+ the following species: Cr(OH)2+, Cr(OH)+2, Cr(OH)3(S), Cr2(OH)4+ 2 , Cr3(OH)4 , and 6+ Cr4(OH)6 in solution (Smith and Martell 1976; Baes and Mesmer 1976).

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III. Toxicity of Cr Of the two most common oxidation states, Cr(VI) is more toxic and soluble whereas Cr(III), is relatively nontoxic and insoluble. In fact, Cr(III) in trace amounts is an essential component of human and animal nutrition (Mertz 1969; Jeejeebhoy et al. 1977) and is an essential activator of insulin associated with glucose metabolism. Chromate is toxic to both plants and animals because it is a strong oxidizing agent, corrosive, and a potential carcinogen (National Research Council 1974). The chromate ion is a class A human carcinogen by inhalation and an acute irritant to living cells (Yassi and Nieboer 1988), and of all the metal carcinogens Cr exhibits properties most nearly consistent with a mutagenic initiation model (Loforth and Ames 1978). A. Toxicity to Plants and Microorganisms Plant growth studies in solution cultures with low levels of Cr (Huffman and Alloway 1973) have indicated that although Cr is essential for human and animal nutrition it might not be an essential component of plant nutrition. Although CrO42− exhibited toxicity effects in plants at concentrations between 18 and 34 mg/kg dry wt (NAS 1974), Skeffington et al. (1976) observed that barley crops could tolerate a Cr-spiked nutrient solution up to 5.0 mg/L, although at 5.0 mg/L a yield reduction of 75% was observed. Generally Cr accumulates in the roots of plants more than in the shoots (Cary et al. 1977a,b; Losi et al. 1994). Mishra et al. (1997) also observed this tendency of Cr accumulation in roots of a paddy crop grown under flooded conditions. Glasshouse plant growth studies using tannery sludge-treated soils have shown toxicity of Cr when sludge was applied at a rate exceeding 3 g/kg, and a significant yield reduction was also observed (Naidu et al. 2000b). In this study the toxicity was attributed to the presence of Cr(VI) in the amended sludge. Similar Cr toxicity has been reported for crops grown in tannery sludgeamended soils from India (Sara Parwin Banu et al. 2000). The adverse effects of Cr(VI) on microorganisms have also been well documented (see article by Kamaludeen et al., this volume). Ross et al. (1981) found that 10–12 mg Cr(VI)/L was inhibitory to most of the soil bacteria in a liquid media, whereas at this level Cr(III) had no effect. Studies using the Microtox test (Beckman Instruments, Carlsbad, CA) have shown that Cr(VI) is approximately twice as toxic as Cr(III) in the Photobacterium phosphoreum-based test at the optimum pH of 6.5 (Qureshi et al. 1984). Two Cr(VI)-resistant bacteria were isolated from the tannery waste-contaminated site at Adelaide and studied for Cr(VI) toxicity. The growth of Arthrobacter sp. was not affected by Cr(VI) up to 50 mg/L whereas 10 mg/L was toxic to Bacillus sp. (Megharaj et al. 1999). B. Toxicity to Animals and Humans Systemic toxicity may occur in both the oxidation states, mainly because of increased absorption of Cr through broken skin that results in renal chromate

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toxicosis, liver failure, and eventually death (Lippmann 2000). Acute exposure of rats to Cr(VI) by various routes of administration affected mainly the liver and kidneys (USEPA 1980). Soluble salts of chromates are also highly toxic when administered parenterally (intravenously), with an LD50 of 10–50 mg/kg, compared to LD50 values of 200–350 and 1500 mg/kg obtained from dermal or oral exposure, respectively (Carson et al. 1986). Conversely, oral administration of Cr(III) compounds is relatively nontoxic. Studies using rats and mice indicated oral LD50 values of 1.87 and 3.25 g/kg for CrCl3 and Cr(NO3)3, respectively (Smyth et al. 1969). The oral LD50 of Na2Cr2O7 in humans is 50 mg/kg (NIOSH 1979). Other effects of Cr(VI) poisoning include gastric distress, olfactory impairment, nosebleeds, liver damage, and yellowing of the tongue and teeth. C. Carcinogenicity and Genotoxicity The carcinogenicity of Cr to animals is well documented (IARC 1980, 1990; Cohen et al. 1993; Costa 1997; Cohen and Costa 1997). These studies also support the hypothesis that some of the most potent carcinogens are the slightly insoluble hexavalent compounds. The Cr(VI) ion is readily taken up into eukaryotic cells by anion-carrying proteins, where it is reduced to Cr(III) by a number of cytoplasmic reducing agents. The reduction of Cr(VI) to Cr(III) causes the generation of oxygen radicals in cells that can produce DNA damage. Additionally the Cr(III) formed can become adducted to the DNA. Recent studies have shown that Cr(VI) is very potent in forming DNA–protein cross-links. This complex typically involves the binding of Cr(III) to the phosphate backbone of DNA and cross-linking to a protein (Lippmann 2000). This cross-linking may lead to increased mutagenicity and is probably more significant in determining the mutagenicity of Cr than the oxidative DNA damage produced by oxygen. Although earlier sections have elucidated the general chemistry of Cr, oxidation states, and toxicity of various species in the environment, the following sections review the chemistry of Cr specifically in relation to soil and water environments and discuss the implications for Cr-contaminated sites.

IV. Chemistry of Soil Cr The concentration of Cr in soil is determined by the amount present in the parent material from which the soil was formed, plus the amounts added through wind, water, and human activities, minus the amounts removed through leaching, surface runoff, volatilization, and phyto-uptake. The literature contains numerous reviews on the chemistry of Cr in soil and water. These reviews describe factors that influence the transformations between the two stable Cr species of Cr(VI) and Cr(III) in soils (Cary 1982; Bartlett and James 1988; Fendorf 1995; Proctor et al. 1997; Naidu and Kookana 2000). The most important of these reactions of Cr in the soil and water environment are summarized in the Cr cycle (Fig. 2).

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Fig. 2. Chromium cycle in soil and water. (Modified from Bartlett and James 1988.)

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A. Chromium Cycle in Soil and Water In soil the predominant species is Cr(III). However, Cr(VI) is also found at significant levels. Because Cr(VI) is a toxic species and reduction reactions are more favored in most soil environments, the starting point for the Cr cycle is Cr(VI). The speciation of chromate in soil solution is pH dependent and dominated by HCrO−4 or CrO2− 4 . Chromium can be temporally removed from soil by adsorption/precipitation reactions, uptake by plants, or leaching from the subsurface layers if soil pH and moisture conditions are favorable. Some of the Cr(VI) even undergoes reduction to Cr(III) by carbon during photosynthesis. In addition, species such as Fe2+ or S2− or products of microbial decomposition of soil organic matter can serve as the direct electron donors. Bartlett and James (1988) referred to this step as “dechromification,” and it is the most important part of the Cr cycle. In terms of decontamination, this step is also important because during this step the toxic form of Cr [Cr(VI)] is reduced to the nontoxic form [Cr(III)]. Reduced Cr(III) is typically bound by a variety of ligands in soil solution that render it insoluble, immobile, and unreactive. However, ligands such as citrate can mobilize Cr(III) and deliver it to MnO2 surfaces where both the organic ligand and Cr become oxidized. MnO2 surfaces are formed by the oxidation of Mn by either atmospheric O2 or soil microbes. Sometimes the organic ligand is recycled because Mn(III) formed by reverse dismutation accepts electrons from the Cr(III) in preference to those from the organic ligand and therefore only Cr is oxidized. When organic ligands are in excessive concentrations they tend to induce reverse dismutation of MnO2 by binding the Mn(III), and this Mn(III)–organic complex may prevent the formation of Cr(VI) or reduce Cr(VI) as fast as it forms, short-circuiting the cycle or speeding up cycling (Bartlett and James 1988). For the present review, most attention was directed toward the first step of the Cr cycle, i.e., partitioning of Cr in soil fractions, soil solution factors influencing adsorption and desorption of Cr(VI), and reduction of Cr(VI) to Cr(III) in soils. In the following sections, the first reactions of the Cr cycle involving both the Cr species are reviewed and the implications for tannery waste-contaminated soil are discussed.

V. Solid-Phase Speciation of Cr in Native Soils Soils derived from serpentines usually contain higher concentrations of Cr and nickel (Ni). Noble and Hughes (1991) used a sequential extraction scheme of 0.5 M KNO3 for 16 hr; deionized water for 2 hr; 0.5 M NaOH for 16 hr; 0.05 M Na2EDTA for 6 hr; and finally 4 M HNO3 for 16 hr at 80 °C to extract various soil fractions of Cr and Ni in serpentine soils from two sites of eastern Transvaal, South Africa. In all nine soils studied, 95% of the total Cr was extractable by HNO3, whereas the general order of other extractants was NaOH > EDTA Ⰷ KNO3 = H2O (Fig. 3).

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Fig. 3. Fractions of Cr in three serpentine soils. (Data source: Noble and Hughes 1991.)

In another study, Xiao-hou et al. (1993) extracted fractions of Cr in 25 typical soils from across China using a different sequential extraction scheme (Table 1). The total Cr in the soils varied from 24 to 110 mg/kg. The majority of Cr (81.5%) was found in the residual fraction and in the crystalline Fe oxide fractions (9.8%). All other fractions combined accounted for less than 10.8% of the total Cr. A. Solid-Phase Speciation of Cr in Sludge and Sludge-Amended Soils Sequential extractions studies are designed primarily for sludge-amended or sludge-contaminated soils rather than native soils where the sequential extraction scheme of Tessier et al. (1979) is typically followed. Sludge high in Cr and Zn was added to a loamy sand soil, and over 2 yr Dudka and Chlopecka (1990) studied the partitioning of Cr in the different soil fractions. The sequential extraction scheme of exchangeable (CaCl2), carbonate (NaOAc), Fe-Mn oxides (NH2OH ⴢ HCl), organics (HNO3 + H2O2), and residual (HF + HClO4) developed by Tessier et al. (1979) was followed. Predominantly Cr was partitioned in the organic fraction of the sludge and sludge-treated soil (Fig. 4). However, in the untreated soil more was partitioned in the residual fraction. Exchangeable and carbonate-bound Cr was negligible compared to total Cr. Zufiaurre et al. (1998) studied the solid-phase speciation of 13 metals includ-

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Table 1. Sequential extraction scheme used to fractionate Cr in 25 Chinese soils. Soil : solution ratio

Equilibrium conditions

Step Fraction

Extractant

1

1M Mg(NO3)2 (pH 7.0)

1:4

Shake 2 hr, 25 °C

1M NH4OAc (pH 5.0) 0.1M NH2OH.HCl (pH 2.0) (a) 30% H2O2 added in two steps (b) 1M NH4OAc (pH 5.0) 0.2M (NH4)2C2O4 + 0.2M H 2C2O 4 0.2M (NH4)2C2O4 + 0.2M ascorbic acid Digest with HF-HClO4

1 : 10 1 : 10

Shake 6 hr, 25 °C Shake 1 hr, 25 °C

1 : 10 1 : 10 1 : 20

(a) Dry by water bath (85 °C) then cool (b) Shake 2 hr, 25 °C Shake 4 hr in dark, 25 °C

1 : 20

Water bath 1 hr, 96 °C

2 3 4

5 6 7

Exchangeable Carbonate Mn oxide bound Organically bound Amorphous Fe oxide Crystalline Fe oxide Residual

Source: Xiao-hou et al. (1993).

Fig. 4. Partitioning of Cr in sludge, sludge-treated, and normal soils. Source: Dudka and Chlopecka 1990.

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ing Cr in sewage sludge from an urban wastewater plant in Spain using the sequential extraction scheme of Tessier et al. (1979). Partitioning of Cr was found mostly in the organic matter fraction (46%) followed by the residual fraction (40%). The average amount of total Cr present was 785 mg/kg and the sludge had an average organic matter content of 48.4%. Similar results were reported by Fernandez et al. (2000) for urban sewage where the sequential extraction scheme of Tessier et al. (1979) was compared with a method recently developed by the Community Bureau of References (BCR) (Quevauviller et al. 1993). These studies clearly indicated that in general Cr is strongly bound to the residual fraction in native soils and predominantly bound to the organic fraction in sludge and sludge-amended soils. Water-soluble and exchangeable Cr fractions were negligible in both native and sludge-amended soils. B. Solid-Phase Speciation of Cr in Industrial Contaminated Soils Few studies to date have examined the fractionation of Cr in industrial contaminated soils. Of the clay loam, loam, and sandy clay loam soils collected from a heavy metal-contaminated site, Wasay et al. (1998) studied the fractionation of Cr only in the clay loam soil. Using the sequential extraction scheme proposed by Fiedler et al. (1994), they found that of the total 832 mg/kg of Cr, 486.9 mg/ kg (58.5%) was bound to the organic soil fraction. Phillips and Chapple (1995) used the sequential extraction scheme of Tessier et al. (1979) to fractionate Cr along with other metals from a soil collected from a former industrial site within inner Brisbane. Past activities at the site involved shipbuilding and a government poisons store. Chromium concentrations in all soil fractions were low, and approximately 80% of the Cr were associated with organic and oxide fractions with negligible concentrations detected in the exchangeable and carbonate fractions. In another study (Maiz et al. 1997), soil samples collected from a polluted mine works, and steel factory and highway emissions were sequentially extracted to find the partitioning of Cr and other metal fractions. A short threestep sequential extraction scheme was compared with the modified Tessier et al. (1979) method and the modified Ure et al. (1993) method. However, because different extracting concentrations were used in all three extraction schemes direct comparison of the results was problematic. However, Cr was found to be predominantly partitioned in the residual fraction of the soils using all three extraction schemes. In all the aforementioned studies, Cr fractionation was investigated simultaneously with several other metals and the concentration of Cr present in the soils was not very high. In addition, no attempt was made to distinguish between Cr(III) and Cr(VI); however, there was good agreement between the sum of various fractions and the total Cr present in the soil. In the following section, sequential extraction studies specifically targeting Cr in highly Cr-contaminated soils are reviewed.

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C. Solid-Phase Speciation in High-Cr-Contaminated Soils Milacic and Stupar (1995) studied the fractionation and oxidation of Cr in tannery waste and sewage sludge-amended soils using the sequential extraction scheme described in Table 2. This sequential extraction scheme was significantly different from the extraction schemes previously used to fractionate Cr in the soils and described in the preceding sections. The use of 0.015 M KH2PO4 for water-soluble Cr is appropriate for targeting the Cr(VI) in Cr-contaminated soils. However, Bartlett and James (1979) showed that a 0.01 M phosphate buffer is more suitable for extracting the exchangeable Cr(VI) from soils. Hence, the use of 0.015 M KH2PO4 for the water-soluble fraction might overestimate the water-soluble fraction by extracting Cr(VI) as well from the adsorbed phase. Fractionation of Cr in natural soils had shown that most of the Cr (80%) was bound to the residual fraction in clay and sandy soils. In peat soils a substantial amount of Cr (20%) was bound to the organic fraction but 70% of the total Cr was still partitioned in the residual fraction. The water-soluble and exchangeable fractions of Cr were very low in all three soils. The fractionation of Cr in tannery sludge showed an entirely different partitioning trend. In tannery sludge 65% of the Cr was bound to oxides and hydroxides and about 20% of the Cr was bound to organic matter and 13% existed as carbonates. Compared to natural soils, substantially more water-soluble (1.5%) and exchangeable (0.5%) Cr was extracted from tannery sludge. Total concentration of Cr was very high in the sludge, about 30,000 mg/kg. The partitioning of Cr in tannery sludge amended clay and sandy soils was similar to that of Cr partitioning in the sludge. However, in the peat soil Cr was bound primarily to the organic fraction (35%), followed by the oxides and hydroxides (25%) and the residual fraction (20%).

Table 2. Sequential extraction scheme to fractionate Cr in soil, tannery sludge, and sludge-amended soils.

Step Fraction

Extractant

1. 2. 3. 4. 5.

Water soluble Exchangeable Organic bound Carbonate bound Oxides and hydroxides

6. 7.

Chromium sulfide Residual

0.015M KH2PO4 1M NH4Cl 0.1M Na4P2O7 1M NaCH3COO, pH 5 0.04M NH2OH.HCl in 25% v/v acetic acid, pH 2 1M HNO3 HNO3 + HClO4 + HF

Source: Milacic and Stupar (1995).

Equilibrium conditions Shake 2 hr Shake 4 hr Two extractions, 4 and 18 hr Two extractions, 4 and 18 hr Two extractions, 90 °C, 5 hr Two extractions, 4 and 18 hr

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In another study Chuan and Liu (1995) using the Tessier et al. (1979) scheme, sequentially extracted Cr from a pig tannery sludge and other tannery plants representative of Taiwan. Ninety-three percent of the 17,000 mg total Cr/kg soil was bound to the Fe-Mn oxides while the remainder was partitioned in the organic and residual fractions and Cr in the exchangeable fraction was negligible. However, speciation of Cr was not carried out in their study. In Cr-contaminated aquifer sediments, Asiakinen and Nikolaidis (1994) studied the partitioning of Cr using a scheme composed of exchangeable [0.01 M KH2PO4/K2HPO4 (1:1) solution at pH 7.2], organic [0.1 M Na4P2O7 at pH 10], Fe or Mn oxide [0.1 M NH2OH ⴢ HCl in 25% acetic acid (v/v), pH 2], and residual [reflux with HNO3 and H2O2]. Of the 346 mg total Cr/g soil in the soil, 25% was found in the exchangeable fraction, 11% was bound to organic matter, 30% was bound to Fe/Mn oxide, and 34% was tightly bound to the residual soil matrix. Thangavel and Naidu (2000) sequentially fractionated Cr, in increments of 10 cm from the surface, from tannery waste contaminated soil down to a depth of 70 cm. They used a modified sequential extraction scheme previously used for serpentine soils by Noble and Hughes (1991). This scheme involved a watersoluble fraction (deionized water, repeated three times), exchangeable (0.5 M KNO3), organic (0.5 M NaOH), remaining organic plus Fe oxide (0.05 M Na2EDTA), and residual (4 M HNO3) fractions. The total Cr concentration ranged from 17,529 to 61,931 mg/kg for the first 10 cm in the three soil profiles. More than 90% of the Cr was partitioned in the residual fraction, and the order of partitioning was Fe oxide > organic > exchangeable = water. Chromium speciation in water-soluble, exchangeable, organic, and Fe oxide fractions showed significant concentrations of Cr(VI) in the water-soluble and exchangeable fractions. This result implies the presence of significant levels of Cr in a bioavailable and mobile form. D. Partitioning and Mobility of Cr The solid-phase speciation studies clearly indicate that at low concentrations of Cr either in natural or contaminated soils, most of the Cr is partitioned in the residual fraction. In highly organic soils a significant portion of Cr is partitioned in the organic fraction and equally partitioned either in the Fe oxide or residual fractions. The concentration of Cr in the water-soluble and exchangeable fractions is very low and indicates low mobility of Cr from these soils. Although in highly contaminated soils Cr is partitioned predominantly in the organic and Fe oxide fractions, a significant amount of Cr existed in watersoluble and exchangeable fractions. In terms of bioavailability these two fractions are important and, given that Cr predominantly exists in two stable oxidation states, speciation of Cr in these fractions requires further attention. For the exchangeable fraction a 10 mM phosphate buffer proves to be the most suitable extractant, but the efficiency of potassium in a K2HPO4/KH2PO4 buffer to extract exchangeable Cr(III) was not tested. The determination of exchangeable Cr(III)

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is necessary because if soil pH conditions are favorable, and in the presence of MnO2, this fraction could become available for oxidation to toxic Cr(VI). Milacic and Stupar (1995) used 1 M NH4Cl for the exchangeable fraction and 0.015 M KH2PO4 for the water-soluble fraction of Cr in soils. However, information on the labile and exchangeable pools of Cr(VI) in contaminated soils is lacking, and such information may be useful in understanding the desorption chemistry of Cr(VI) and is essential for the development of a suitable remediation strategy for contaminated soils.

VI. Factors Affecting the Soil Solution Chemistry of Cr The presence of Cr in soil solution is determined by many soil factors including pH, redox potential, type of clay minerals, organic matter content, nature of the adsorbed phase, soil solution composition, and, importantly, the source of the pollution that determines the initial form of Cr added. Each of these factors is discussed in detail. A. Soil Solution pH The phase diagram depicting the thermodynamic stability of aqueous Cr species over a range of pe (redox) and pH is presented in Fig. 5 (Calder 1988). Cr(III) is the most stable form under reduced (low redox) conditions and is present as a cationic species with the first or second hydrolysis products dominating at pH values from 4 to 8. Under oxidized (high redox) conditions, Cr(VI) is often thermodynamically the most stable species observed in the pH range 2 to 14. However, Cr(VI) is readily reduced to Cr(III) in the presence of reducing agents, and thereafter Cr(III) can form highly soluble organic complexes under acidic conditions (Calder 1988). B. Chromium (III) in Soil Solution The dominant species of Cr(III) in water at pH between 6 and 8 is Cr(OH)+2. Under acidic conditions, CrOH2+ and Cr3+ dominate, and Cr(OH)03 and Cr(OH)−4 dominate under alkaline conditions. The governing reactions are the following: Cr3+ + H2O ⇔ CrOH2+ + H+ CrOH2+ + H2O ⇔ Cr(OH)+2 + H+ Cr(OH)+2 + H2O ⇔ Cr(OH)03 + H+ Cr(OH)03 + H2O ⇔ Cr(OH)−4 + H+ Cr(III) forms complexes with many organic ligands as well as fluoride, ammonia, cyanide, thiocyanate, oxalate, and sulfate (Beas and Mesmer 1976). Bartlett and Kimble (1976a) studied the behavior of 10 mmol Cr(III)/kg added to a spodic horizon, a Spodosol Ap horizon, and a clay Alfisol A horizon having pH in the range 4–4.8. They also studied the effects of increasing pH levels, adjusted with CaCO3, and extraction with NH4OAc (pH 4.8), 1 M HCl,

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Fig. 5. Thermodynamic stability of aqueous Cr species over a range of Pe and pH encountered in the surface environment. (From Calder 1988.)

1 M NaF, and 0.12 M Na4P2O7. These results indicated that only HCl and Na4P2O7 extracted significant portions of the added Cr(III) and that NH4OAc and NaF only removed very small fractions of Cr. In two Spodosols, as characterized by the NH4OAc and NaF extracts, Cr(III) decreased as the pH was increased by addition of CaCO3, and the effect was similar in clay soils for NH4OAc but not for NaF. It was inferred that as the pH of the studied soils was increased, the extractability of Cr(III) decreased. In the same study, Bartlett and Kimble (1976a) considered the relationship between pH and solubility for Cr(III) in a silty loam soil and solution by raising the pH via addition of either CaCO3 or NaOH. As solution pH was raised above 4, the solubility of Cr(III) decreased, with complete precipitation occurring at pH 5.5. The precipitate was believed to be polymeric with Cr ions in six coordination complexes with water and hydroxyl groups (Mertz 1969). Grove and Ellis (1980) also observed a similar relationship between Cr(III) solubility and

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pH. When Cr(III) precipitates as amorphous Cr(OH)3, the expected concentration of aqueous Cr(III) is less than 0.05 mg/L (Rai et al. 1989). Cifuentes et al. (1996) added different amounts of Cr(III) as CrCl3.6 H2O to a slightly acidic (pH 6.2) soil and two alkaline soils (pH 7.9 and 8.3) and found that the concentration of Cr(III) in solution increased with the total Cr(III) added; this was the result of a low equilibrium soil pH (<3.8) at the end of the experiment. At these low pHs MINTEQA2, a computer program model for metal speciation in soil solution (Allison et al. 1991), predicted no precipitation of Cr(III) and that the aqueous speciation would be dominated by Cr3+ and Cr(OH)2+. In our studies in long-term tannery waste-contaminated soils, soluble Cr(III) was detected in water extracts from the subsurface horizon (>50 cm) having an acidic pH of 4, and enhanced release of Cr(III) was also observed when extracted with 0.01 M CaCl2 (Avudainayagam et al. 2001). Although the presence of Cr(III) in soil solution is highly pH dependent, studies in long-term tannery-contaminated soil have indicated that the effect of aging favors stronger binding of Cr(III) to adsorption sites in the soil. C. Solubility of Organically Complexed Cr(III) Chromium is bound either in inorganic or organic pools and can become available for further sorption or conversion to more toxic Cr(VI) by oxidation reactions on dissolution. The extremely low solubility of inorganic forms such as Cr(OH)3 at pH > 5.5 suggests that leaching of Cr(III) to groundwater or oxidation of Cr(III) to Cr(VI) from inorganic complexes is rare in most soils. However, pollution problems could arise with organically complexed Cr(III) as these complexes are generally more soluble and therefore more mobile than the inorganically chromium complexes in soils. James and Bartlett (1983a) studied the relationship between pH, organic ligands, and the behavior of Cr(III). The organic ligands included many different naturally occurring carboxylic acids found in soils and roots [e.g., citric acid and diethylenetriamino pentaacetic acid (DTPA)]. They found that Cr(III) in citric acid remained soluble up to pH 7–7.5, whereas Cr(III) in water (no organic ligand added) was precipitated between pH 4.5 and 5.5. The solubility trend of Cr(III) in DTPA was similar to that observed in citric acid except that Cr(III) in DTPA was less soluble than Cr citrate at pH between 6 and 7. They also studied the complexation of Cr(III) by organic polymers such as fulvic acid and water-soluble organic matter leached from an air-dried soil sample. Fulvic acid effectively complexed Cr(III) and prevented its precipitation up to approximately pH 7.5, if the ratio of fulvate COOH/Cr = 1. James and Bartlett (1983a) also studied the long-term solubility of inorganic and organically complexed Cr(III) in soils at different pHs. In their incubation experiments, they observed that water-soluble Cr(III) decreased rapidly during the first week regardless of whether it was added as Cr citrate or as CrCl3. However, during the next 7 weeks, soluble Cr(III) in the citrate treatments decreased more than inorganic CrCl3-amended soils. Indeed, soluble Cr(III) could

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be detected even after 1 yr of incubation, when applied as Cr citrate, regardless of the soil pH. Fractionation of Cr in tannery sludge revealed significant concentrations of water-soluble Cr (450 mg/kg) (Milacic and Stupar 1995) and of this watersoluble Cr only a small fraction (0.9 mg/kg) was speciated as Cr(VI). Thus, most of the water-soluble Cr existed as Cr(III), but details on the actual pH of the tannery sludge were not reported. However, Naidu et al. (2000a) reported a pH range of 7.7–9.4 for the tannery sludge collected from a tannery in South Australia. Hence, it may be concluded that even at alkaline pH, organically complexed Cr(III) could be available in soil solution. D. Adsorption and Desorption of Cr(III) Cr(III) sorption and desoprtion kinetics are important determinants of the availability of soluble forms of Cr for further oxidation to toxic Cr(VI) in soils. Cr(III) has been shown to be sorbed strongly on to soil minerals, to be bound to soil organic matter, and to form mineral precipitates (Bartlett and Kimble 1976a; Cary et al. 1977a; Rai et al. 1987, 1989; Bartlett and James 1988; Palmer and Wittbrodt 1991; Losi et al. 1994). Sorption of Cr(III) decreases when other inorganic cations or dissolved organic ligands are present in solution. Fendorf et al. (1994) and Fendorf and Sparks (1994) have studied the mechanism of Cr(III) sorption on silica using extended absorption fine structure spectroscopy and found that Cr(III) formed a monodentate surface complex on silica. Arnfalk et al. (1996) studied Cd, Hg, Pb, Cr(III), and Cr(VI) retention on 14 different types of minerals and soil materials considering both pH dependency and other soil physicochemical parameters. The results verified the importance of geochemical parameters of soils such as organic content, type of clay mineral, presence of complexing ions, and redox potential for controlling metal uptake. Montmorillonite (in bentonite and smectite) showed the highest retention of Cd, Cr(III), and Pb among all minerals and soil materials, whereas illite and kaolinite showed lower retention than the soils. The clay mineral montmorillonite showed highest retention because it had the highest surface activity (Kashef 1986). The difficulty in displacing Cr from smectite indicates that the Cr is bonded specifically because if Cr was held through outer sphere complexes, the smallest hydroxy polymers would be readily displaced by Ca (Dubbin and Goh 1995). Drljaca et al. (1992) found that, while the montmorillonite was still wet, the adsorbed Cr could be easily exchanged with other cations but, upon drying, Cr becomes virtually nonexchangeable. These authors suggested that as the interlayer region collapsed due to loss of water, Cr came into close contact with the siloxane surface, allowing inner sphere complexes to form. Transition metals with a 3d3 electronic configuration must overcome a large energy barrier to form intermediates, and ligand displacement reactions of Cr(III) complexes are extremely slow (Cotton and Wilkinson 1988). Cr(III) is held strongly, likely through covalent bonds, and its displacement is extremely difficult through sim-

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ple exchange reactions. However, the potential for Cr(III) to be oxidized to the more toxic Cr(VI) form is of some concern because of the instability of bonding under strong oxidizing conditions. Both adsorption and precipitation reactions and both specific and nonspecific reactions are possible for the retention of Cr(III) in soils. However, organically complexed Cr(III) could be available in soil solution even at high soil pH for oxidation to toxic Cr(VI) in soils.

VII. Oxidation of Cr(III) to Cr(VI) Oxidation of Cr(III) to Cr(VI) represents a significant environmental hazard because a relatively nontoxic species is transformed into a more toxic one. Manganese oxides are the only naturally occurring oxidant of Cr(III) and oxidation of Cr(III), in the presence of MnO2, was first observed by Bartlett and James (1979). They noted that Cr(VI) was present in the effluents of most soils reacted with Cr(III). Even the manganese oxide with the highest zero point charge and the most crystalline structure, pyrolusite, is an effective oxidant of Cr(III) (Eary and Rai 1987; Saleh et al. 1989). Adsorption of Cr(III) by Mn oxides is possibly the first step in its oxidation by Mn. In soils, manganese oxides typically accumulate on the surface of clay and iron oxides at relatively high redox potentials. McKenzie (1977) noted that Mn minerals tend to have large surface areas and high negative charge at all but extremely acidic pH. These properties are associated with high adsorptive capacities, particularly for heavy metals. Cr(III) can be oxidized to Cr(VI) in the presence of Mn4+ where Mn4+ acts as the oxidizing agent and is reduced to Mn2+, as shown by the equation: 2+ + 2Cr3+ + 3MnO2 + 2H2O ⇔ 2CrO2− 4 + 3Mn + 4H

A. Effect of pH Bartlett and James (1979) studied the influence of pH from 3 to 10 on Cr(III) oxidation in a soil having 1200 mg/kg of 1 M HCl extractable manganese. In this study, the soil solution was shaken for 0.5 min of every 2 min for 18 hr. They observed that at pH 3.2 all the Cr(III) was oxidized to Cr(VI) and that about one-quarter of the Cr(VI) formed was adsorbed by soil. The oxidation and adsorption decreased with increasing pH. The percentage of Cr(III) oxidation to Cr(VI) decreased from 100% to 30% at pH 6 and stayed constant at 30% up to pH 9. They used a large solution/soil ratio (2000 : 1) to control the change of pH caused by addition of CrCl3 to the soil. B. Solubility of Cr(III) and Oxidation As previously discussed, solubility and availability of Cr(III) in soil solution are critical for the oxidation of Cr(III) to Cr(VI) in soils, although at soil pH > 5.5 the solubility of Cr decreases due to its precipitation as Cr(OH)3. Complexation

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of Cr(III) with some of the low molecular weight organic acids such as citrate and gallic acid increases its solubility and mobility even at higher pH, thereby facilitating its oxidation. James and Bartlett (1983b) compared the oxidizing tendencies of four forms of Cr(III) added to a field moist soil incubated for 15 d. The four forms were freshly precipitated Cr(OH)3, Cr citrate, aged Cr(OH)3, and aged Cr(OH)3 with citrate. The maximum Cr(VI) levels observed decreased in the order freshly precipitated Cr(OH)3 > Cr citrate > aged Cr(OH)3 in citrate > aged Cr(OH)3. Milacic and Stupar (1995) studied the oxidation of Cr(III) in tannery waste amended to three soil types. Their fractionation study showed that after 5 mon 1.1% of the total Cr added was oxidized in clay, 0.45% in sand, and only 0.03% in peat soil. The degree of Cr(III) oxidation was found to be proportional to the concentration of manganese (IV) oxides and water-soluble Cr(III) in the soils. They also observed a decrease in the concentration of water-soluble Cr and Cr(VI) on continuance of the experiment because Cr was redistributed to more sparingly soluble fractions.

VIII. Factors Affecting Soil Solution Cr(VI) The pe–pH stability phase diagram (Fig. 5) illustrates the dominant aqueous Cr(VI) species present under equilibrium conditions. Under oxidizing conditions and from pH 2 to pH 14, anionic Cr (VI) dominates. The HCrO−4 species of Cr(VI) dominates up to pH 6, and above pH 6 it dissociates to form CrO2− 4 in dilute aqueous systems. In the soil solution the concentration of Cr(VI) is dependent on soil pH, redox potential, adsorption–desorption processes, competing ions in soil solution, reduction reactions, and the nature of Cr(VI) from the source of pollution. A. Effect of pH Soil solution pH plays a significant role in the availability and mobility of Cr(VI) in soil. Soil pH affects the magnitude and sign of charges on soil colloids, especially on organic matter and on Fe(III), Al(III), and Mn(III,IV) oxides (McKenzie 1977; Parfitt 1978). Because Cr(VI) is anionic it is not retained appreciably on the negatively charged colloids of soils or sediments. However, Cr(VI) is attracted to positively charged surfaces, such as iron, manganese and aluminium oxides, and hydroxides (Stumm and Morgan 1981). Therefore, binding of Cr(VI) to soil surfaces will depend on soil mineralogy and on the relationship between soil pH and the zero point of charge of the colloids involved. Bloomfield and Pruden (1980), in a study on the leaching behavior of Cr(VI) in a near neutral (pH = 6.7) and an acidic (pH = 4.2) soil from the Sawyers field of the Rothamsted Experimental Farm, found that in neutral soils, after 3 wk of contact, most of the Cr(VI) was removed unchanged by leaching water. Although reduction of Cr(VI) was a major factor with the acidic soils, about half

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the added Cr(VI) persisted in sorbed or labile forms. These observations demonstrated the potential for Cr(VI) leaching in alkaline soils. Naidu et al. (2000b) and Mahimairajah et al. (2000) presented evidence for Cr(VI) in groundwater from highly Cr-contaminated alkaline soils. From their batch leaching experiments on a chromite ore processing wastecontaminated soil from New Jersey, Weng et al. (1994) reported that at pH < 2.5, the amount of Cr(VI) leached was below their detection limit (<0.01 mg/L). At pH between 2.5 and 4.5, the amount of Cr(VI) leached sharply increased and reached a maximum at pH 4.5. The amount of Cr(VI) leached remained constant between pH 4.5 and 12. They attributed this to adsorption of Cr(VI) onto the soil under acidic conditions. The pHzpc of the studied soil was 6.8, and electrostatic interactions favor anionic chromate species adsorption when the solution pH is less than pHzpc; electrostatic chromate adsorption onto Cr soil is not favorable at a high solution pH. Davis and Olsen (1995) studied the adsorption of Cr(VI) onto two sandy soils by mixing the soil with groundwater spiked with various concentrations of Cr(VI). They observed low Kd (Freundlich adsorption constant) values of 0.16 and 0.21 L/kg for the two soils and attributed this result to competition of SO42− present in the groundwater for the adsorption sites, a process that would enhance the Cr(VI) release in these neutral pH soils. Although Henderson (1994) only measured an average KdCr of 0.31 L/kg for Trinity sand aquifer material of pH 7.5, Stollenwerk and Grove (1985) reported a KdCr(VI) of 52 L/kg at pH 6.8 for alluvial sands. Sandy material has a greater preponderance of positively charged surfaces over the pH range 5–7.5, compared to clay (isoelectric point pH 2–3.5), and this results in greater affinity for CrO2− 4 and higher KdCr(VI) on sand than for clay. Davis and Olsen (1995) compiled the published data on Cr(VI) sorption data demonstrating the effect of pH and soil type on KdCr(VI) and observed a linear relationship between Kd and pH for sand and clays (Fig. 6). James (1994) studied the leaching behavior of Cr(VI) in chromite ore processing residue-contaminated soils by extracting with distilled water, 1.0 mM KNO3 in 0.5 mM K2SO4 (salt solution, pH 7), and 1.0 mM HNO3 in 0.5 mM H2SO4 (acid solution, pH 2.7). James observed no difference between distilled water, salt solution, and dilute acid in extraction of the Cr(VI) from two soils. Despite the fact that the repeated leachings of the soil with acid gradually lowered the pH of the extracts, Cr(VI) concentrations were unchanged from water and salt solutions. He attributed this to the presence of a solid phase containing sparingly soluble Cr(VI) for a gradual and steady release of Cr(VI) from the contaminated soils. Depending on the source of chromate contamination and the nature of adsorbed phase, the solubility of Cr(VI) could be independent of soil solution pH. Studies in our laboratory demonstrated a similar gradual and steady leaching of Cr(VI) from surface alkaline soil contaminated with tannery waste at Mount Barker, South Australia (Naidu et al. 2000a; Avudainayagam et al. 2001).

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LEGEND Sand Kaolinite Montmorillonite

Sheppard et al. 1987

Kd-Cr (L/kg)

120

80 Griffin et al. 1977

Sheppard et al. 1987

Stollenwerk and Grove, 1985

40 Griffin et al. 1977

0

This study Griffin et al. 1977

Masschelyn, et al. 1992

4

6

Henderson, 1994

8

pH Fig. 6. Sorption data demonstrating the effect of pH and soil type on KdCr(VI) sand (dashed line) and clays (dotted line). (From Davis and Olsen 1995.)

B. Adsorption of Cr(VI) Adsorption of Cr(VI) by soils and standard minerals has been studied extensively. Modeling efforts of Cr(VI) sorption on hydrous oxides of Fe and Al and in soils suggest that Cr(VI) forms an outer sphere complex on these surfaces (Benjamin and Bloom 1981; Zachara et al. 1987, 1989; Ainsworth et al. 1989). In contrast to results of surface complexation modeling, competitive ion displacement studies have shown that Cr(VI) is retained much more strongly than anions such as Cl− or SO2− 4 and that its retention strength approaches that of phosphate (Bartlett and Kimble 1976b; Griffin et al. 1977). To ascertain whether Cr(VI) forms an inner sphere or outer sphere complex, Fendorf (1995) employed X-ray absorption fluorescence spectroscopy (XAFS) to identify the Cr(VI) surface complex on Al and Fe hydrous oxide. He found that Cr(VI) forms an inner sphere complex on goethite, with both bidentate and monodentate complexes (Fig. 7). Based on these results, one would therefore expect that Cr(VI) also forms an inner sphere complex on Al hydrous oxides. However, Aide and Cummings (1997) proposed an outer sphere complex of Cr(VI) adsorption on Al oxyhydroxide (boehmite). They demonstrated that Cr(VI) adsorption on boehmite is pH dependent and this pH dependency is − strongly related to the protonation of CrO2− 4 to form HCrO4. They also observed suppression of Cr(VI) adsorption by dilute concentrations of o-phosphate competing for similar substrate sites. Therefore, it is possible that both specific and nonspecific adsorption of Cr(VI) can be observed in soils where pH and soil

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Fig. 7. Surface complexation structure of Cr(VI) on goethite. Note the interatomic distances and atom types coordinating Cr(VI). (From Fendorf 1995.)

solution composition may have greater influence on the solubility and mobility of Cr(VI) in soils.

IX. Contaminated Soil Environment and Mobility of Cr(VI) The nature of the contaminated soil environment depends largely on the source of contamination, whether from dumping of industrial solid or liquid wastes. The nature and composition of the waste would therefore influence the solubility and precipitation of Cr(VI) disposed along with the waste to soil. Because the tanning industry generates approximately 50,000 tons of solid waste per year from chromium-mediated tanning in India and surrounding countries (Ramasamy and Naidu 2000), and between 80,000 and 100,000 (wet) tons of sludge annually in the United States (Ho et al. 1982), this source of Cr is expected to dominate any discussions of Cr contamination of the environment. A. Tannery Sludge Composition Chrome tanning produces two types of sludge, hair-burn and chrome. The hairburn or dehairing sludge results from the initial stages in the tanning process, when salt-cured hides are mechanically scraped and soaked in Ca(OH)2 to remove hair. This residual sludge contains higher concentrations of organic carbon, total N, Ca2+, Mg2+, and Na+, but relatively small amounts of Cr. After

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dehairing, the hides are cleansed and soaked in a NaCl–H2SO4 solution, then bathed for several hours in a basic Cr(SO4)3 tanning solution. The hides are then neutralized with NaHCO3. This last stage yields the chrome sludge, with high Cr concentration along with high concentrations of organic carbon, total N, Ca2+, Mg2+, and Na+ (Driess 1986). However, the composition of sludge varies from place to place and industry to industry and depends on the choice of chemicals, raw material, pretreatment methods for the wastewater employed, and above all the quality of water used for the tanning processes. For a comparison, the composition of tannery sludge from India, South Australia, and the United States is presented in Table 3. B. Soil Solution Composition and Mobility of Cr(VI) In terms of ionic activity and solubility of Cr(VI) in soil solution, the presence 2− in tannery-contaminated of high concentrations of Ca2+, Na+, PO3− 4 , and SO4 soils have a significant influence on the mobility of Cr(VI) in such soils. James and Bartlett (1983b) studied the effect of sulfate and phosphate on the adsorption of Cr(VI) to two A and two B horizon soils and found that phosphate has a greater effect than sulfate in decreasing Cr(VI) adsorption by the soils. This finding was not surprising given that phosphate is a stronger ligand than sulfate. Aoki and Munemori (1982) showed that adsorption of chromate by iron oxide

Table 3. Chemical composition of tannery chrome sludge from different countries.

Composition of tannery sludge pH EC (dS m−1) Total organic carbon (%) Cr (%) CaCO3 (%) Ca (%) Na (%) Fe (mg kg−1) Mn (mg kg−1) Cu (mg kg−1) N (mg kg−1) P (mg kg−1) K (mg kg−1) NR: not reported. aFrom Naidu et al. (2000a). bFrom Lollar (1982). cFrom Sara Parwin Banu et al. (2000).

South Australiaa

USAb

South Indiac

7.71 3.5 11 3.04 11.7 4.7 4.7 1.1 1.3 66.5 NR 4032 3549

NR NR 15.4 3.86 NR 1.8 0.7 NR NR NR 31000 NR 1900

7.8 3.5 9.6 0.08 NR 0.8 NR NR NR NR 4300 5100 40

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was increased in the presence of cadmium, lead, copper, and zinc divalent cations. Increased adsorption may be attributed to specific adsorption of these cationic metals, which would reduce the net surface negative charge of colloid particles. Zachara et al. (1987) studied the chromate adsorption on amorphous iron oxyhydroxide in the presence of major groundwater ions. They found that in paired solute systems (e.g., CrO2− 4 –H2CO3), anionic cosolutes markedly reduce adsorption through a combination of competitive and electrostatic effects, CrO2− 4 but observed no appreciable influence of cations (K+, Mg2+, and Ca2+). However, definitive studies on the effect of dominant cations and anions present in the tannery waste on Cr(VI) mobility in a long-term real contaminated soil is lacking.

X. Remediation of Chromate-Contaminated Soil and Water A number of treatment technologies are available for soils that are contaminated with high levels of Cr(VI) other than conventional remediation by excavation and off-site disposal. A wide array of treatment technologies are also available to remove chromate from water. However, the selection and application of the specific type of treatment technology would depend upon many soil and water factors specific to the contaminated site, some of which are reviewed in this section. A. Remediation Strategies for Contaminated Soil The technologies applicable to a particular Cr site depend on the cleanup goals, the forms of Cr present, the volume of soil to be remediated, and the physiochemical properties of the Cr-containing soils. Although Cr (III) does not pose a significant groundwater or surface runoff hazard, Cr(VI) is highly mobile in the soil and water environment, is acutely toxic at moderate doses, and is classified as a known respiratory carcinogen in humans (USEPA 1996a). Most of the developed countries have proposed distinctly different risk-based criteria for Cr(III) and Cr(VI) in soils. The threshold levels of both forms of Cr permissible in soils recommended by some developed countries are listed in Table 4. Region III of the U.S. Environmental Protection Agency (USEPA) has established a risk based residential soil cleanup level of Cr(III) of 78,000 mg/ kg and Cr(VI) of 390 mg/kg based on an ingestion pathway (USEPA 1996b). The U.S. Environmental Protection Agency has also designated a soil screening level of 270 mg/kg Cr(VI) for potential human exposure by inhalation (USEPA 1996c). The National Environmental Protection Council of Australia (NEPM 1999) is currently developing a new soil investigation level for various contaminants including both Cr(VI) and Cr(III). The new soil investigation levels of Cr are categorized into health investigation and ecological investigation levels. Another cleanup criterion often used for sites is the Toxicity Characteristic Leaching Procedure (TCLP), and the limit for total Cr is 5 mg/L (40 CFR

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Table 4. Threshold levels of Cr(III) and Cr(VI) permissible in soils of some developed countries (mg/kg). Australiaa 50g

New Zealandb

UKc

Canadad

USEPAe

600h 10i

6002 253

2502 83

15002 —

Germanyf 2002 —

aANZECC/NHMRC (1992); bMinistry of Environment/Ministry of Health (1995); cUK Department of Environment (1987); dCCME (1990); eUSEPA (1993); fMarshall (1994); gTotal Cr; hCr(III); iCr(VI).

261.24; as quoted by Higgins et al. 1997). This procedure measures leachable Cr concentrations, predominantly Cr(VI). Therefore, in terms of setting cleanup standards for Cr, the highly mobile and toxic Cr(VI) is given more importance than the immobile Cr(III) in the soil and water environment. Soil Remediation Technologies. Apart from removing the Cr(VI)-containing soils from a site, other available technologies consist of (1) extracting the Cr(VI) from soils, (2) immobilizing the Cr so that it will not leach, as defined by the TCLP test, and (3) irreversibly reducing the Cr(VI) in soils to the Cr(III) valence state. The list of treatment methods under each category and their advantages and disadvantages are presented in Table 5. At sites where the extent of contamination is large (hectares) and total Cr concentrations are high, in situ treatment methods are the most suitable options. If a considerable amount of total Cr(VI) was observed, then irreversible reduction of Cr(VI) to Cr(III) by using an appropriate reducing agent would be an effective and economical approach of remediation. The use of the land after decontamination has been completed should be considered in selecting suitable in situ treatment methods. Irreversible Reduction of Cr(VI) to Cr(III). The literature indicates that extensive research has been directed toward the reduction of toxic and mobile Cr(VI) to relatively nontoxic and immobile Cr(III) in both soil and water (James and Bartlett 1983b; Eary and Rai 1988; Powell et al. 1995; Wittbrodt and Palmer 1995, 1996a,b; Fendorf and Li 1996; Buerge and Hug 1997; Blowes et al. 1997; Patterson et al. 1997; Seaman et al. 1999). Hexavalent Cr can be reduced to Cr(III) in soils by redox reactions with aqueous inorganic species, electron transfers at mineral surfaces, reaction with nonhumic organic substances such as carbohydrates and proteins, or reduction by soil humic substances (Palmer and Wittbrodt 1991). However, the selection of a suitable reducing agent is a prerequisite in achieving irreversible reduction of Cr(VI) to Cr(III) in contaminated soils. The attainment of irreversible reduction is a prerequisite because of the possibility of reoxidation of Cr(III) back to Cr(VI) by manganese oxides under favorable soil conditions.

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Table 5. Remediation technologies for chromate-contaminated soils. Technology

Advantage

Disadvantage

Removal of Cr (VI) Excavation and offsite disposal

Appropriate for small volumes of soil, quick to implement, uses readily available equipment, completely removes Cr

May cause Cr(VI) to become airborne and a health hazard; may invoke land disposal restrictions; can be expensive especially for deep material; soil may need treatment, adding to the cost

Soil washing

May greatly reduce the volume of contaminated material that requires treatment

May cause Cr(VI) to become airborne and a health hazard during excavation; Cr is transferred from soil to water source; not appropriate for soils with strongly bound Cr(VI)

Soil flushing

In situ technology, does not require excavation Demonstrated technology for remediating Cr(VI) spills

Generates contaminated water; not appropriate for soils with strongly bound Cr(VI)

Relatively inexpensive Demonstrated technology (USEPA 1995)

May cause Cr(VI) to become airborne and a health hazard during excavation; does not destroy contaminants; generates increased volume of solidified mass that requires disposal; Cr(VI) may need to be reduced to Cr(III); not appropriate for sites where total Cr levels must be decreased; may require soils dewatering

In situ technology, does not require excavation Applicable to sites with high water table

Cr(VI) may first need to be reduced to Cr(III); not appropriate for sites where total Cr levels must be reduced

Immobilization of Cr(VI) Solidification/ stabilization ex situ

Solidification/ stabilization

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Table 5. (Continued). Technology Vitrification

Reduction of Cr(VI) to Cr(III) Chemical reduction

Biological reduction

Advantage

Disadvantage

May be performed in situ Reduce and immobilize Cr(VI) May produce marketable byproduct

High energy requirements, may require soil dewatering, may require significant amounts of additives

May be performed in situ Applicable to beneath buildings and other substances

Not appropriate for sites where total Cr levels must be decreased; other compounds in soil may consume reducing chemicals and so process may require significant amounts of reducing agents Not appropriate for sites where total Cr levels must be decreased; requires control of pH, oxygen, nutrients, etc. Process may be slow to achieve Cr(VI) reduction

In situ technology, does not require excavation, applicable to sites with high water table Applicable to sites where Cr(VI) continues to leach from soil Applicable to beneath buildings and other structures

Several effective Cr(VI) reducing agents have been tried including the ferrous iron (Eary and Rai 1988), sulfide (Pettine et al. 1994), and organic complexes (Elovitz and Fish 1995; Deng and Stone 1996; Lay and Levina 1996). Additionally, direct enzymatic reduction of chromate has been demonstrated by a number of soil microorganisms (Ishibashi et al. 1990; Llovera et al. 1993). Bartlett and James (1976b) found that soils high in organic matter reduced Cr(VI) to Cr(III) rapidly over a pH range of 4.4–7.6. However, James and Bartlett (1983a,b) demonstrated that Cr(III) forms soluble chelated complexes with organic reductants and observed that even the organically complexed Cr(III) can be oxidized to Cr(VI) in soils. Aqueous ferrous iron and sulfide can also effectively reduce Cr(VI) to Cr(III) (USEPA 1980; Pettine et al. 1994; Eary and Rai 1988). The reduction of Cr(VI) to Cr(III) by ferrous iron is very rapid and is described by the following general equation: Cr(VI) (aq) + 3Fe(II) (aq) ⇔ Cr(III) (aq) + 3Fe(III) (aq)

(1)

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Eary and Rai (1988) studied the reduction of aqueous Cr(VI) by Fe(II) and observed that the products of the reduction reaction (1) precipitated as hydroxide solids in slightly acidic to alkaline solutions depending on the solution concentrations: xCr(III) + (1 − x)Fe(III) + 3H2O ⇔ (CrxFe1−x) (OH)3(s) + 3H+

(2)

where x can vary between 0 and 1. Cr(III) hydroxide solids, including Cr(OH)3(s) and the solid solution (Cr,Fe)(OH)3(s), have been observed to precipitate rapidly in moderately acidic to alkaline solutions and are expected to be important solubility-controlling solids for dissolved Cr(III) over a wide range of pH (Rai et al. 1987; Sass and Rai 1987). They observed that the solubility of this solid limits Cr(III) concentrations over the pH range of approximately 5–11 to less than the drinking water standard of 0.05 mg/L Cr. Therefore, based on the published reports on aqueous Cr(VI) reduction by reducing agents, it is obvious that use of ferrous iron may be the best option to achieve irreversible reduction of Cr(VI) to Cr(III). However, very limited studies are found in the literature reporting the use of ferrous iron for reducing Cr(VI) to Cr(III) in soil and especially in contaminated soils. James (1994) studied the reduction of water-soluble and exchangeable Cr(VI) in alkaline soils enriched with chromite ore processing residue using various reducing agents including ferrous iron. He observed that, compared to Mn2+, steel wool, dried hardwood tree leaf litter, and ferrous iron reduced all soluble and exchangeable Cr(VI) in the high-hex soils studied. However, he noted that ferrous iron acidified the soil more than the Mn2+ treatments. Seaman et al. (1999) studied in situ reduction of Cr(VI) within a coarse-textured, oxide-coated soil and aquifer system using Fe(II) solutions; they found a decrease in the concentrations of Cr(VI) in the batch and column effluent with increasing Fe(II). However, dissolved Cr, presumably Cr(III), exceeded regulatory limits due to the low pH (3.0) induced by oxidation and hydrolysis of Fe(II). To date there are few published reports on Cr(VI) reduction in tannery wastecontaminated soils and also no studies on reoxidation of reduced Cr(III), which is surprising given that the Cr contamination arises chiefly from tannery wastes dumping. B. Remediation of Cr(VI)-Contaminated Water Chromium is used in a number of industries that generate large volumes of Cr(VI)-contaminated wastewater, and some of these industries may also be generating wastewater containing Cr(VI) as the dissolved Cr species, namely, Cr plating and chemical milling and etching industries (USEPA 1979). Apart from the generation of large volumes of Cr(VI) dissolved in industrial wastewater, indiscriminate disposal of Cr-rich wastewater to land has resulted in contamination of groundwater with Cr(VI). Groundwater contamination by Cr(VI) is more common in industrialized areas in developing countries, and

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large plumes of Cr(VI)-contaminated groundwater in shallow sand and gravel aquifers have been well documented (Permutter and Lieber 1970; Deutsch 1972; Wiley 1983; Stollenwerk and Grove 1985; French et al. 1985). Disposal of tannery effluent and sludge onto the land has resulted in severe contamination of soil and groundwater quality by Cr in the Vellore district of Tamil Nadu, India (Mahimairaja et al. 2000). Therefore, to safeguard dwindling natural resources, it is necessary to treat contaminated groundwater and industrial wastewater to remove toxic Cr(VI). Water Treatment Methods to Remove Cr(VI). Chemical precipitation is one of the most common techniques used for the removal of heavy metals from wastewater. The chemicals most frequently used for precipitation of metals are lime, caustic soda, and sodium carbonate. Although most heavy metals can be precipitated readily by pH adjustment, Cr(VI) is highly soluble and does not precipitate out of solution at any pH. Consequently, treatment for Cr(VI) usually consists of a two-stage process: first, the reduction of Cr(VI) to Cr(III) and, second, the precipitation of Cr(III). Reducing agents commonly employed are gaseous sulfur dioxide or a solution of sodium bisulfate (Lanouette 1977). Other suitable reducing agents include sodium sulfite, sodium hydrosulfite, and ferrous sulfite (Kraljik 1975). However, there are no published reports available on direct removal of Cr(VI) from water. Tokunaga et al. (1999) demonstrated that lanthanum (La) salts can effectively remove As(III) and As(V) in water. They showed that La was effective over a wide range of pH and was able to remove As to meet effluent and drinking water standards with a much smaller dose of La than would be required by other conventional precipitants. Hexavalent Cr, being an oxy-anion like As, could behave similarly and precipitate from water by lanthanum salts, but this has yet to be tested.

XI. Future Research Needs This review highlights a number of areas in which the current knowledge of Cr chemistry, especially in tannery waste-contaminated sites, is lacking that necessary to develop cost-effective remediation solutions for such sites. More effort needs to be directed toward the speciation of Cr in long-term contaminated sites rather than the laboratory-based studies where native soils are spiked with fresh salts and then aged for brief periods. This need reflects the fact that the chemistry of Cr in long-term contaminated sites varies markedly from that observed in laboratory studies. Although it is now commonly accepted that the toxicity of Cr would depend strongly on the speciation of total Cr between Cr(III) and Cr(VI), few studies have considered the speciation of Cr at all. In terms of bioavailability, knowledge of the distribution of these two oxidation states, which dominate environmental Cr chemistry, is vital for determining relevant end points and for guiding legislators in assigning realistic health and ecological investigation levels.

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Clearly, much work needs to be directed toward the speciation of Cr in highly contaminated soils. Related to this, information on the labile and exchangeable pools of Cr(VI) in contaminated soils is also lacking, and an understanding of the desorption chemistry of Cr(VI) is essential for the development of suitable remediation strategies for contaminated soils.

Summary Worldwide chromium contamination of soils has arisen predominantly from the common practice of land-based disposal of tannery wastes under the assumption that the dominant species in the tannery waste would be the thermodynamically stable Cr(III) species. However, significant levels of toxic Cr(VI) recently detected in surface water and groundwater in India, China, Australia, and elsewhere raise critical questions relating to current disposal criteria for Cr-containing wastes. It now appears that despite the thermodynamic stability of Cr(III), the presence of certain naturally occurring minerals, especially Mn oxides, can enhance oxidation of Cr(III) to Cr(VI) in the soil environment. This factor is of public concern because at high pH, Cr(VI) is bioavailable, and it is this form that is highly mobile and therefore poses the greatest risk of groundwater contamination. A review of the current literature indicates that extensive research has been performed on the speciation of Cr in soil, the effect of pH on soil solution concentrations of Cr(III) and Cr(VI), soil adsorption phenomenon of Cr species, redox reactions, and transformation of Cr(III) and Cr(VI) together with remediation strategies to decontaminate Cr-contaminated soils. Most of the studies were conducted using an uncontaminated soil artificially spiked with Cr, and very limited research has been conducted in the contaminated soil environment. Furthermore, studies on tannery waste contaminated soils are limited, and obviously a serious gap of knowledge exists in understanding the influence of long-term tannery waste contamination on Cr behavior in soil.

Acknowledgments Financial support from the John Allwright Scholarship of the Australian Centre for International Agricultural Research (ACIAR) to S. Avudainayagam and Remediation of Contaminated Environments Program, CSIRO Land and Water is acknowledged. We thank Dr. N. Sethunathan for critical reading and suggestions.

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Tessier A, Campbell PGC, Bisson M (1979) Sequential extraction procedure for the speciation of particulate trace metals. Anal Chem 51:845–851. Thangavel P, Naidu R (2000) Fate and behaviour of chromium at the long-term tannery waste-contaminated site near Adelaide. In: Naidu R, Willet IR, Mahimairaja S, Kookana R, Ramasamy K (eds) Towards Better Management of Soils Contaminated with Tannery Waste. ACIAR (Australian Centre for International Agricultural Research) Proceedings, no 88. Canberra, Australia, pp 71–74. Thorstensen TC (1976) Practical Leather Technology. Krieger, New York. Tokunaga S, Yokoyama S, Wasay SA (1999) Removal of arsenic(III) and arsenic(V) ions from aqueous solutions with lanthanum(III) salt and comparison with aluminium (III), calcium(II) and Iron(III) salts. Water Environ Res 71:299–306. UK Department of Environment (1987) ICRCL—Guidance on the Assessment and Redevelopment of Contaminated Land. Guidance Note 59/83, 2nd Ed. x, x. Ure AM, Quevauviller P, Muntau H, Griepink B (1993) Speciation of heavy metals in soils and sediments. An account of the improvement and harmonization of extraction techniques undertaken under the auspices of the BCR of the Commission of the European Communities. Int J Environ Anal Chem 51:135–151. USEPA (U.S. Environmental Protection Agency) (1979) Water-related environmental fate of 129 priority pollutants. EPA 440/4-79-029A. Office of Water Planning and Standards, Washington, DC. USEPA (U.S. Environmental Protection Agency) (1980) Health assessment document for chromium. USEPA, Environmental Criteria and Assessment Office, Research Triangle Park, NC. USEPA (U.S. Environmental Protection Agency) (1993) Standards for the use or disposal of sewage sludge. Fed Reg 58:210–238. USEPA (1995) Innovative Treatment Technologies: Annual Status Report. 7th Ed. EPA542-R-95-008. USEPA, Washington, DC. USEPA (1996a) Region III risk-based concentration table, January–June 1996. Memorandum from RL Smith, Office of RCRA, Technical and Program Support Branch (3HW70), Washington, DC. USEPA (1996b) Integrated Risk Information System (IRIS). Reference dose for oral exposure to Cr (III) and Cr (VI). USEPA, Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office, Cincinnati, OH. USEPA (1996c) Soil screening guidance: technical background document. EPA/540/ R-95/128. Office of Solid Waste and Emergency Response, Washington, DC, p A-5. Wasay SA, Barrington S, Tokunaga S (1998) Retention form of heavy metals in three polluted soils. J Soil Contam 7:103–119. Weng CH, Huang CP, Allen HE, Cheng AH-D, Sanders PF (1994) Chromium leaching behaviour in soil derived from chromite ore processing waste. Sci Total Environ 154: 71–86. Whitten KW, Gailey KD (1984) General Chemistry, 2nd Ed. Saunders, Philadelphia. Wiberg KB (1965) Oxidation of chromic acid and chromyl compounds. In: Wieberg KB (ed) Oxidation in Organic Chemistry. Academic Press, New York. Wiley KG (1983) Hydrogeological investigation examining the identification and cleanup of chromium contaminated groundwater in Richland Township, Kalamazoo County, Michigan. MS Thesis, Wright University, xxxxx. Wittbrodt PR, Palmer CD (1995) Reduction of Cr(VI) in the presence of excess soil fulvic acid. Environ Sci Technol 29:255–263. Wittbrodt PR, Palmer CD (1996a) Effect of temperature, ionic strength, background elec-

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trolytes and Fe(III) on the reduction of hexavalent chromium by soil humic substances. Environ Sci Technol 30:2470–2477. Wittbrodt PR, Palmer CD (1996b) Reduction of Cr(VI) by soil humic acids. Eur J Soil Sci 47:151–162. Xiao-hou S, Guang-xi X, Wen-xing Y (1993) Distribution of chemical forms for Co, Cr, Ni and V in typical soils of China. Pedosphere 3:289–298. Yassi A, Nieboer E (1988) Carcinogenicity of chromium compounds. In: Niragu JO, Nieboer E (eds) Chromium in Natural and Human Environments. Wiley, New York, pp 443–495. Zachara JM, Girvin DC, Schmidt RC, Resh CT (1987) Chromate adsorption on amorphous iron hydroxide in the presence of major groundwater ions. Environ Sci Technol 21:589–594. Zachara JM, Ainsworth CC, Cowan CC, Resch CT (1989) Adsorption of chromate by subsurface soil horizons. Soil Sci Soc Am J 53:418–428. Zufiaurre R, Olivar A, Chamorro P, Nerin C, Callizo A (1998) Speciation of metals in sewage sludge for agricultural use. Analyst 23:255–259. Manuscript received September 30, accepted October 7, 2002.

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 Springer-Verlag 2003

Rev Environ Contam Toxicol 178:93–164

Chromium–Microorganism Interactions in Soils: Remediation Implications Sara P.B. Kamaludeen, Mallavarapu Megharaj, Albert L. Juhasz, Nabrattil Sethunathan, and Ravi Naidu Contents I. Introduction .......................................................................................................... A. Forms of Chromium ....................................................................................... B. Sources of Chromium in Soil ......................................................................... C. Chromium Transformations in Soil ................................................................ II. Physicochemical Factors Governing Chromium Transformations in Soil ........ A. Soil Physical Factors ...................................................................................... B. Soil pH ............................................................................................................ C. Organic Matter ................................................................................................ D. Iron .................................................................................................................. E. Manganese ....................................................................................................... III. Microbiological Factors Governing Chromium Transformations in Soil .......... A. Resistance or Tolerance to Cr(VI) ................................................................. B. Direct Cr(VI) Reduction ................................................................................. C. Indirect Reduction ........................................................................................... D. Biotic–Abiotic Coupling in Mn Oxide-Mediated Oxidation of Chromium(III) ............................................................................................ IV. Implications of Chromium Transformations on Microorganisms and their Activities .............................................................................................. A. Microorganisms .............................................................................................. B. Effects on Soil Microbial Community ........................................................... C. Effect on Soil Microbial Processes and Activities ........................................ V. Remediation of Chromium-Contaminated Water and Soils ............................... A. Remediation Technologies for Wastewater and Solutions ............................ B. Remediation Technologies for Chromium Wastes in Soils .......................... C. Bioremediation ................................................................................................ D. Applicability of Phytostabilization to Cr-Contaminated Soil ........................ VI. Challenges ............................................................................................................ Summary ....................................................................................................................

94 95 95 97 97 97 97 97 98 99 103 103 104 119 121 126 126 130 132 141 141 143 144 147 148 148

Communicated by G.W. Ware. S.P.B. Kamaludeen The University of Adelaide, Department of Soil and Water, Waite Campus, Glen Osmond, SA 5064, Australia and Tamil Nadu Agricultural University, Trichy Campus, Trichy, Tamil Nadu, India. M. Megharaj ( ), A.L. Juhasz, N. Sethunathan, R. Naidu, (formerly CSIRO Land and Water, Adelaide), Australian Centre for Environmental Assessment and Remediation, University of South Australia, Mawson Lakes Campus, Mawson Lakes, SA 5095, Australia.

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Acknowledgments ...................................................................................................... 149 References .................................................................................................................. 149

I. Introduction The increasing urbanization and human population worldwide has generated an ever-increasing amount of inorganic and organic wastes of domestic and industrial origin. Such wastes have generally been disposed onto land for centuries, relying on the soil’s capacity to decontaminate waste materials by biological and physicochemical means and render them harmless by adsorption or precipitation of potential pollutants in the wastes (Martin and Parkin 1985). Continued and excessive loading of such wastes beyond the soil’s capacity as a sink, however, has led to disastrous consequences to soil and water resources worldwide. Industrial disposal of tannery wastes onto soil had been a common practice before enactment of stringent regulations in many countries including Australia. One of the major problems encountered with disposal of tannery wastes is the presence of chromium (Cr), a heavy metal, in the waste. Chromium is widely used in the metallurgic, refractory, chemical, and tannery industries. Chrome plating, the deposition of metallic Cr, imparts a refractory nature to materials, rendering them resistant to microbial attack and flexible over extended periods of time (Barnhart 1997). More than 170,000 tonnes of Cr waste are discharged to the environment annually as a consequence of industrial and manufacturing activities (Gadd and White 1993). Of the total Cr used in the processing of leather, 40% is retained in the sludge. Generally, tannery wastes contain Cr(III) as the predominant species of Cr together with high concentrations of organic carbon derived from the animal hides. Because of the thermodynamic stability of relatively low toxic Cr(III), disposal of tannery sludge onto land and into water bodies has led to increased Cr levels, reaching as high as 30,000 mg kg−1 or more in contaminated soils (Naidu et al. 2000b). Chromium is considered as one of the priority pollutants in the United States by the U.S. Environment Protection Agency (USEPA), and in many other countries, primarily because the soluble Cr species, Cr(VI), is a respiratory carcinogen when inhaled and a mutagen as a result of its strong oxidizing nature (USEPA 1996a). In contaminated soils, in the absence of reducing agents, Cr(VI) is soluble in alkaline environments, posing a threat to surface and groundwater quality because it is more readily transported. For these reasons, regulatory authorities monitoring the contaminated sites have placed considerable emphasis on remediation and rehabilitation of Cr-polluted soils. The techniques for remediation of Cr-contaminated soils are based mostly on transforming the toxic Cr(VI) to nontoxic Cr(III) species and immobilization of Cr(III) by precipitation (Higgins et al. 1997). However, there is still a danger that detoxified forms may later revert to toxic species because of the changes in soil properties, different farming techniques, or climatic variables. Such rapid reversible transformations of Cr in soil have further complicated the task of determining whether Cr-bearing waste or waste-contaminated soil is hazardous as recorded in the Federal

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Register in 1991 (James et al. 1997) and of setting the guidelines for remediation. This review highlights (i) Cr transformations in soil, (ii) abiotic and biotic factors governing Cr transformations in soil, (iii) effect of Cr on soil microbial activities, and (iv) remediation of Cr-contaminated soils with special emphasis on bioremediation techniques. A. Forms of Chromium Chromium, categorized as a heavy metal, is the 24th element in the periodic table and is prevalent in nature as the 17th most abundant element on earth. Chromite (FeOCr2O3) is the only major commercial product. Chromium occurs in oxidation states ranging from Cr(II) to Cr(VI) (Avudainayagam 2002; see also Avudainayagam et al., this volume), but Cr(III) and Cr(VI) are the two stable oxidation states of Cr in soil and water environments. Cr(VI), the form used in many industrial applications and also formed in the environment through oxidation of Cr(III), is relatively more water soluble, bioavailable, reactive, and toxic than Cr(III). Chromium compounds are mutagenic (Venitt and Levy 1974) and teratogenic (Bauthio 1992). Generally, Cr(III) has a low toxicity, whereas Cr(VI), inhaled through dust, is recognized by the International Agency for Research on Cancer and by the U.S. Toxicology Program as a pulmonary carcinogen (Barceloux 1999). In addition, Cr(VI) causes irritation to and corrosion of the skin and respiratory tract in humans and an allergic contact dermatitis known as eczema. B. Sources of Chromium in Soil Chromium was first discovered in crocoite in 1798 and named after the bright color of Cr compounds. In the soil environment, it is derived from both natural and anthropogenic sources. Natural Sources Chromium is found preferentially in ultrabasic and basic rocks, feldspar materials in particular. On average, the earth crust contains 3,700 mg Cr kg−1, most of which resides in the core and mantle (Nriagu 1988). The Cr concentration of the inner core is about 12,100 mg kg−1 (Liu 1982). Most of the chrome ores are located in three major places: deposits of the bushved complex of South Africa, the great dyke in Zimbabwe, and the kemi intrusion of Finland (DeYoung et al. 1984). The global production of Cr annually amounts to 107 tons. Anthropogenic Sources Chromium compounds are widely used in many industries, especially in metallurgical, refractory, and chemical manufacture. These industries use low-grade chromite ores for numerous applications including pigment manufacture, metal finishing, corrosion inhibition, organic synthesis, leather tanning, and wood preservation (USEPA 1988).

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The annual Cr consumption in different industries is given in Fig. 1. The leather industry alone accounts for 40% of the worldwide Cr usage. Large-scale disposal of tannery wastes has significantly contributed to Cr contamination in soils and water worldwide. Most of the Cr reaches the soil by improper disposal of industrial wastes, spills, or faulty storage containers (USEPA 1984). Approximately 50,000 tonnes of Cr-rich solid wastes are disposed onto land annually from tannery industries alone. The long-term disposal of tannery wastes has led to extensive contamination of agricultural soil and groundwater in several countries, including Australia, China, India, Bangladesh, Nepal, Pakistan, Spain, and Brazil. About 50,000 hectares of land have been rendered barren by this activity in India and Bangladesh alone (ACIAR 2000). A long-term contaminated site at Mount Barker near Adelaide (South Australia) revealed the presence of total Cr at concentrations as high as 70,000 to 100,000 mg kg−1 soil even 20 years after cessation of tannery waste disposal (Naidu et al. 2000a,b).

Refractory 3%

Others 10%

Pigments 15%

Wood preservation 15%

Leather 40%

Metal finishing 17%

Fig. 1. Chromium consumption in different industries.

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C. Chromium Transformations in Soil In soils, Cr(III) and Cr(VI) are the more common forms of Cr. Cr(VI) is soluble, mobile, and hence easily contaminates both groundwater and soil. In contrast, Cr(III) is sparingly soluble, less mobile than Cr(VI), relatively stable in the environment (McGrath and Cegarra 1992), and becomes more inert over time as a result of its precipitation. Following addition of Cr-rich wastes to soil, Cr undergoes rapid transformations and attains a dynamic equilibrium between Cr(III) ⇔ Cr(VI) through a combination of physical, chemical, and biological processes. The major processes governing Cr transformations in soils include adsorption or desorption, redox conditions, and precipitation or dissolution (Nieboer and Jusys 1988).

II. Physicochemical Factors Governing Chromium Transformations in Soil Transformations of Cr in complex and dynamic soil systems are governed by several physical, chemical, and biological factors. A. Soil Physical Factors The influence of soil physical factors in relation to Cr transformations is not documented in detail, except for a few reports. The adsorption and desorption processes are governed largely by the bulk and particle density of the soils and the type and amount of clay and organic matter content. In peat soils, Cr(III) is the more prevalent species because of its binding to organic and mineral fractions. However, in sandy and clay soils, Cr(III) tends to oxidize because of its solubility in easily extractable fractions of soil solution (Milacic and Stupar 1995). Chromium behaves both as anion and cation depending on its speciation. The chemical factors dominate and play a major role in Cr transformations in soil. Some of the important chemical and biological factors are discussed next. B. Soil pH Soil pH plays a significant role in controlling the dynamics of Cr redox reactions (Bartlett and James 1988; Losi et al. 1994c). Soil pH along with redox potential (Fig. 2) determine the nature of Cr prevalent in the soil (Rai et al. 1989). Low pH favors the formation of stable cationic Cr(III) species whereas at higher pH, especially in alkaline soils, anionic Cr(VI) formation is favored. At low pH, Cr(III) precipitates or tightly binds to a variety of ligands such as hydroxyls, humates, and phosphates present in soil. C. Organic Matter Humic substances and organic matter play a major role in the reduction of Cr(VI) to Cr(III). Organic matter during its oxidation reduces Cr(VI) (Bartlett and Kimble 1976). Citric and fulvic acids and water-soluble extracts of air-dried

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Direct oxidation

Indirect oxidation

Cr III

Mn(II)/Fe(II)

Cr oxidisers Mn/Fe cycle

?

Cr VI Mn(IV)/Fe(III) Fig. 2. Possible direct and indirect Cr oxidation reactions in soil.

soils form soluble complexes with Cr(III) (James and Bartlett 1983a). Organically complexed Cr(III) may remain soluble whereas metal ions are quickly adsorbed and precipitated. Complexed Cr(III) hydroxides differ in their solubility. Freshly precipitated Cr(III) such as CrCl3 or Cr(OH)3 is highly soluble compared to aged precipitates and hence becomes amenable to oxidation. Cr oxidation occurs in the following order: freshly precipitated Cr(OH)3 > Cr citrate > aged Cr(OH)3 in citrate > aged Cr(OH)3 (James and Bartlett 1983b). The solubility of these forms of Cr determines their access to microorganisms and the extent of Cr oxidation. In soils that are low in organic matter incorporation of organicrich wastes has been recommended to promote the reduction of Cr(VI) (Bartlett and Kimble 1976). D. Iron Chromium transformations are influenced by other ionic species released during changes in pH. For example, at low pH, the presence of ferrous [Fe(II)] iron increased the rate of Cr reduction (Palmer and Wittbrodt 1991). The reduction of Cr(VI) to Cr(III) in the aqueous (aq) phase was rapid, as described by the following reaction:. Cr(VI) (aq) + 3Fe(II) (aq)

> Cr(III) (aq) + 3 Fe(III) (aq)

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At pH values greater than 4.0, brown precipitates were observed (Pettine et al. 1994), hypothetically via the following reaction: xCr(III) + (1 − x) Fe(III) + 3H2

> (CrxFe1−x) (OH)3 (s) + 3H+

where x varied between 0 and 1. The precipitate formed, presumably Cr0.25 Fe0.75(OH)3, was not conclusively identified (Patterson et al. 1997). At low pH, the addition of small amounts of iron alone can increase the rate of Cr reduction. The ratio of Fe(II) oxidized was proportional to the amount of Cr reduced (approximately 1.0) (Weng et al. 1996). Two mechanisms have been proposed to explain the chemical reduction of Cr by Fe. First, humic substances convert Fe(II) to Fe(III) which in turn reduces Cr. Second, Fe(II) can form FeCrO4, which complexes with chromate. Amorphous iron sulfide minerals such as mackinawite (FS1−x) have the potential to reduce large quantities of Cr(VI) (85%–100%) and form stable CrFe(OH)3 solids (Patterson et al. 1997). Partially oxidized iron was also effective and widely used for subsurface remediation. In subsoils, a combination of Fe(II), organic matter and low pH (4.2–4.3) govern the reduction of Cr(VI) (Powell et al. 1995). E. Manganese Manganese oxides have been implicated in the chemical oxidation of Cr(III) in the soil environment (Kim et al. 2002). Bartlett and James (1979) were the first to report Mn-mediated oxidation of Cr in soils with a pH above 5.0, provided the soil was fresh and moist. The amount of Cr oxidized was proportional to the amount of Mn reduced (exchangeable) and also to the amount of Mn reducible by hydroquinone. The minimum amount of MnO2 necessary for complete oxidation of Cr in soil is not known. Knowledge of the energetics of the reaction in relation to the availability of oxidizable Cr(III) would be desirable. For optimum oxidation of Cr(III) by manganese oxides, the surface of the latter must be relatively free of specifically adsorbed Mn(II) and other heavy metals. The Cr(III) approaches a receptive surface, is quickly oxidized to the anionic form and then repelled by the like negative charges. The rate and amount of Cr(III) oxidized is dependent on the soil type, pH, nature of Cr(III) present in soil, and mineralogy and quantity of Mn oxides available for oxidation. pH Manganese-mediated Cr oxidation is dependent on soil pH and is normally observed in any soil with a pH of 5.0 and above. Up to pH 5.5, Cr(III) oxidation was increased by naturally occurring δ-MnO2 (Fendorf and Zasoski 1992). However, Cr oxidation by β-MnO2 increased with decreasing pH. Oxidation of Cr(III) occurs not via surface-catalyzed reaction with dissolved O2, but by direct reaction with synthetic β-MnO2. The extent of Cr(III) oxidation at lower pH is limited by the strong adsorption of anionic Cr(VI), thereby inhibiting contact between active oxidizing sites on β-MnO2 (Eary and Rai 1987).

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Chromium (III) oxidation by naturally occurring δ-MnO2 was suppressed as pH and Cr(III) concentration increased simultaneously. The reaction products, Mn(II) or Cr(VI), were not limiting for further oxidation. At pH > 3.5, Cr(III) induced alteration in Mn oxide surface, thus limiting the extent of oxidation. However, the oxidation was also dependent on Cr(III) concentration, pH, initial surface area, and ionic strength (Fendorf and Zasoski 1992). Oxidation of Cr(III) by a MnO2 preparation proceeded rapidly at pH 5.5 and 7.5 with identical rates, slowly at pH 3, and very slowly at pH 1 (Amacher and Baker 1982). In addition to pH-dependent charge characteristics of oxide minerals, Mn oxide surfaces might also have permanent negative charges as substitution of Mn(II) and Mn(III) for Mn(IV) occurs during their oxidation. Because of the rapid redox transformations and specific adsorption continually taking place on manganese oxide surfaces, these charges will be temporary. Exchangeable Cr(III) is not found in soils with pHs greater than 4.5 or 5. As the pH of a 1 µM Cr(III) solution was lowered, amount of oxidation by a dilute soil suspension ranged from 20% at pH 7.5 to 100% at pH 3.2. In some systems, the effects of pH on charge characteristics and surface behavior of manganese oxides can mask the effects of pH on Cr speciation and solubility. Soil Type Studies conducted in whole soils to observe the Cr oxidation have shown that the extent of Cr(III) oxidation varied with clay content and the proportion of waste materials containing Cr(III) (Bartlett 1985, 1986). Land disposal of Cr and the associated health effects have been fully reviewed by Chaney et al. (1981). Behavior of Cr(VI) in organic waste is similar to slowrelease nitrate fertilizers, and chromium in sludge could release low levels of Cr(VI) over a period of years. Also, the Cr associated with high molecular weight ligands is not readily oxidized (Amacher and Baker 1982). These authors reported that sludge-borne Cr was not oxidized even over a 4-yr period; however, moist samples showed phosphate-extractable Cr(VI). Cr(III) The oxidation of Cr(III) is directly related to its concentration in soils. Oxidation is also dependent on the moisture content of soils. For instance, James and Bartlett (1983a) observed that soils incubated under moist conditions released 0–41 µmol Cr(VI) L−1 after 4 yr of incubation. In 1-m2 field plots containing 1000 mg Cr kg−1, 0.04% of the Cr was leached as Cr(VI). A Possible Chromium Oxidation Score (PCOS) was calculated based on the four important factors of waste oxidation potential (WOP), soil oxidation potential (SOP), soil reduction potential (SRP), and soil-waste pH modification value (PMV) as per the following formula: PCOS = WOP + SOP + SRP + PMV (Chaney et al. 1996). Higher values indicated a greater possibility for Cr oxidation in soils. This approach can be used as a quick screening tool to determine the oxidative ability of soils. Recently, the nature of Cr precipitates on goethite (a crystalline ferric oxide) and silicon dioxide was studied using scanning force microscopy. This study demonstrated a new concept in that the ability of the

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precipitating phase to fit the crystal lattice structure of the material on which it is precipitating can cause a significant change in the form and stability of Cr precipitates. Chromium precipitated on goethite spread evenly on the surface and had a low extractability with oxalate compared to Cr precipitating on silica as clumps of Cr(OH)3 that were extracted more rapidly with oxalate (Fendorf et al. 1996) Thus, studies of the chemical nature of Cr precipitated in different soils might also add to our understanding of how soil chemistry can influence the potential oxidation of Cr(III) (Chaney et al. 1996). Mineralogy of Mn Oxides Manganese oxides have a high adsorptive capacity for metal ions, thus potentially providing local surface environments in soil. The oxidation of Cr(III) to Cr(VI) occurs after adsorption of Cr to the Mn, with simultaneous formation of Mn(II). The overall reaction follows: Cr(OH)+2 + 1.5 MnO2 + H2O

> HCrO−4 + 1.5 Mn2+ + H2

The rate of transformation is, however, governed by the mineralogy of Mn, soil pH, and the form and solubility of Cr(III) in soil (Bartlett and James 1988; Milacic and Stupar 1995). Dissolved oxygen has no effect on Cr(III) oxidation by Mn oxides. There was a proportional increase in oxidation of Cr(III) as the surface of Mn oxide increased. Pyrolusite (β-MnO2) in solution oxidized 15.6% of the spiked Cr(III) (96 µM) at pH 3.0 after 400 h (Eary and Rai 1987). Acidic pH favored the dissolution of Mn oxides and enhanced the oxidation of Cr(III). As the pH increased from 3 to 4.3, the rate of Cr oxidation decreased. β-MnO2(s) + 2H+

> Mn2+ + H2O + 1/2 O2 (aq)

Birnessite (δ-MnO2), the predominant Mn oxide in soils, effected more than 90% Cr(III) oxidation. The reaction was also faster (24 hr) with birnessite than with pyrosulite (β-MnO2). pH had a similar effect on oxidation, i.e., acidic pH favored more oxidation. There was no difference in rate of oxidation until pH 4.0, whereas beyond this pH there was a decrease up to pH 5.2. Cr oxidation was also observed at pH 6.3, 8.3, and 10.1 in spite of the low solubility of Cr(III) at these pH levels. Increases in Cr(III) concentration (200–800 µM) retarded its oxidation. The overall reaction for δ-MnO2 was given as Cr3+ + 1.5 δ-MnO2 + H2O

> HCrO−4 + 1.5 Mn2+ + H+

At pH 5.0, CrOH2+ + 1.5 δ-MnO2

> HCrO−4 + 1.5 Mn2+

Stoichiometry reactions indicate that 1.5 moles of Mn(II) was produced for every mole of Cr(VI) formed. Generally, oxidation of Cr(III) decreased with increasing pH and Cr(III) concentration (Eary and Rai 1987), and several hypotheses have been postulated to explain the inhibition.

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During the Cr(III) and MnO2 reaction, only a portion of MnO2 is available for oxidation. There is evidence that Mn(II) and Cr(VI) inhibit Cr(III) oxidation. The more Mn(II) formed, the lower the chance for Cr(III) to react due to poisoning of the surface. In acidic pH, the MnO2 has a negative charge and Mn(II) and Cr(III) compete for binding sites. However, with an increase in binding of Mn(II) the negative charge on MnO2 surfaces is lowered; this in turn increases the pH of the surface because of −OH ions, and Mn(II) becomes autoxidized by atmospheric oxygen. When Mn(II) becomes autoxidized, the surface becomes more negative and is available for further Cr(III) oxidation. However, Fendorf et al. (1993) showed that oxidation of Cr(III) was not inhibited by the addition of Mn(II) to the system. In contrast, Eary and Rai (1987) proposed that Cr(VI) formed was a limiting factor in Cr(III) oxidation. The initial Cr(III) oxidation was instantaneous, and Cr(VI) formed becomes strongly bound to β-MnO2 surfaces, especially at acidic pHs. Moreover, this adsorption also retarded the dissolution of MnO2. The dissolution of MnO2 occurred more in Cr-free solution than with Cr in the solution (Eary and Rai 1987). At neutral and alkaline pHs, Cr(OH)3 precipitates. The difference in mechanism between δ- and β-MnO2 was attributed mainly to the differences in zero point charges (pH 2.3 for δ-MnO2 and 7.3 for β-MnO2). Formation of MnOOH as the intermediate at higher pHs in β-MnO2 can also lead to decreased oxidation of Cr(III). Cr(OH)2+ + 3 β-MnO2(s) + 3 H2O

> HCrO−4 + 3 MnOOH(s) + 3 H+

γ-MnOOH is formed during the reduction or oxidation of Mn oxides as an intermediate (Johnson and Xyla 1991) The oxidation kinetics of Cr(III) to Cr(VI) on the surface of manganite (γ-MnOOH) is a function of Cr(III) and manganite concentration, pH, ionic strength, and temperature. The reaction is first order with respect to the manganite adsorption density and also Cr(III) concentration up to a critical adsorption density (0.2 µmol m−2). Above this concentration, the reaction is inhibited. The reaction is independent of pH and ionic strength. Considering the chemical speciation, the overall reaction at pH 4.5 can be written as Cr(OH)2+ + 3 MnOOH

> HCrO−4 + 3Mn2+ + 3OH

The oxidation rate of Cr(III) is 10–10,000 times faster with γ-MnOOH than with other Mn oxides. Such fast rates of Cr(III) oxidation by γ-MnOOH contribute significantly in the cycling of Cr in natural water systems. Nakayama et al. (1981) found that in seawater only 10% of a 10−5 M solution of Cr(III) was oxidized with 30 mg γ-MnOOH L−1 in 100 hr. The low oxidation note may be linked to the organic substances in natural waters. Experiments with salicylate clearly indicated the inhibition of Cr(III) oxidation reaction with γ-MnOOH by organic ligands.

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III. Microbiological Factors Governing Chromium Transformations in Soil Biological factors also play a major role in the transformations (Cr reduction in particular) and mobilization of Cr in soils. A wide variety of heterotrophic microorganisms is involved in the reduction of Cr(VI) to Cr(III), aerobically or anaerobically depending on the organism, in both soil and water environments (Lovley and Phillips 1994). Evidence suggests that Cr(VI)-reducing microorganisms are ubiquitous in soils and can enhance the detoxification of Cr(VI) under ideal physicochemical conditions (Turick et al. 1996). Chromium(VI)-tolerant and -sensitive bacteria, with ability to transform Cr(VI) to Cr(III), occur widely in diverse ecological conditions: water, sediments, and soil (Losi et al. 1994a). The important microbial processes influencing Cr transformations are considered in this section. A. Resistance or Tolerance to Cr(VI) A variety of mechanisms have been implicated in the adaptation, tolerance, and resistance of microorganisms to a metal pollutant: extracellular precipitation, decreased uptake (resulting from an efflux system, blockage in uptake, or both), and enzymatic reduction [Cr(VI) ⇒ Cr(III)] or oxidation [As(III) ⇒ As(V)] to a less toxic form. Chromate tolerance in microorganisms has probably evolved during their long-term exposure to naturally occurring or anthropogenic sources of Cr or other metals, for instance, Cu (Badar et al. 2000), in the environment. Interestingly, even soils with no previous history of Cr contamination can harbor bacterial populations resistant to Cr(VI). Thus, Bader et al. (1999) found that both contaminated soil with a high Cr level of 12,400 mg/kg soil and two uncontaminated soils harbored aerobic bacterial populations resistant to Cr(VI). However, bacterial populations resistant to Cr(VI) at concentrations as high as 500 µg/mL could be isolated only from the uncontaminated soils and not from the contaminated soil samples. In contrast, fungal colonies resistant to Cr(VI) at concentrations as high as 1000 µg/mL were routinely isolated from both uncontaminated and contaminated soils. Evidently populations resistant to Cr(VI) have evolved in soils, not necessarily related to their previous exposure to Cr. Contaminated soil with a high Cr content of 12,400 mg/kg soil contained much a smaller and less diverse microbial population than that in the uncontaminated soils. For effective bioremediation, Cr(VI) resistance through decreased uptake alone may not be desirable; decreased uptake coupled with its ability to reduce Cr(VI) to less toxic Cr(III) would be particularly useful. Chromate resistance is either plasmid mediated (Summers and Jacoby 1978; Bopp et al. 1983; Ohtake et al. 1987; Silver and Misra 1988) or caused by chromosomal mutations (Cervantes and Silver 1992). Chromosomal and plasmid determinants operate by different mechanisms, as evident by their additive effects. Mutations in bacterial cells caused by DNA damage by Cr(VI) have been reported. Plasmids conferring chromate resistance have been reported in Strepto-

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myces lactis (Efstathiou and McKay 1977), Pseudomonas aeruginosa (Summers and Jacoby 1978; Cervantes and Ohtake 1988), P. fluorescens (Bopp et al. 1983; Ohtake et al. 1987), P. ambigua (Horitsu et al. 1983), and Alcaligenes eutrophus (Nies et al. 1989). Ohtake et al. (1987) found that decreased uptake of chromate, responsible for chromate resistance in P. fluorescens, was plasmid borne. Accumulation of chromate in the resistant strain was 2.2 times less than that in a plasmid-cured sensitive strain. Chromate-resistant strains of P. fluorescens (Ohtake et al. 1987) and A. eutrophus (Nies et al. 1989), however, transported sulfate to the same extent as the plasmid-cured chromate-sensitive strains. Chromate resistance in P. fluorescens was not related to sulfate transport. Currently, eight proteins of the chromate resistance (Chr) family have been fully sequenced (Nies et al. 1998). Among these, Chr proteins from Pseudomonas aeruginosa (Chr Pae) (Cervantes et al. 1990) and Alcaligenes eutrophus (Chr Aeu) have been functionally characterized for their role in chromate resistance. Chr proteins conferred decreased uptake of chromate in chromate-resistant P. aeruginosa (Cervantes et al. 1990) and A. eutrophus (Nies et al. 1990). Recent evidence suggests that decreased accumulation of chromate in P. aeruginosa, conferred by chromate-resistant protein (ChrA), was associated with its active efflux from the cytoplasm driven by the membrane potential (Alvarez et al. 1999). Everted membrane vesicles of P. aeruginosa harboring the Chr A plasmid accumulated four times more chromate than did the vesicles from plasmid-cured cells. There is no evidence yet, however, for chromate efflux as a mechanism for reduced accumulation of Cr in chromate-resistant A. eutrophus. Alcaligenes strain CH34 has determinants encoding inducible resistance to chromate whereas A. alcaligenes AE104 (plasmid-free derivative of CH34) is chromate sensitive. A lux-coupled chromate biosensor, A. eutrophus AE104 (pEBZ141), carrying chrulux transcriptional fusion, developed using a cloned part of plasmid pMOL28 carrying a chromate-resistant determinant from strain CH34, was specific for chromate (Peitzsch et al. 1998). Data from this study on interactions between chromate resistance and chromate reduction and between sulfate concentration and chromate induction showed that chromate resistance was best induced by chromate and bichromate whereas induction by Cr(III) was 10 times lower. Sulfate starvation increased uptake and reduction of chromate in strain 104. B. Direct Cr(VI) Reduction In soils, microbial Cr reduction may occur directly or indirectly. In the direct mode, Cr is taken up by the microbes and then enzymatically reduced (Komori et al. 1990b; Losi et al. 1994c; Lovley and Coates 1997), while in the indirect mode, products (reduction or oxidation) of microbial decomposition in the soil such as H2S mediate the reduction of Cr(VI) (DeFilippi and Lupton 1992). Direct microbial reduction of Cr(VI) was first reported in the 1970s (Lebedeva and Lyalikova 1979; Romanenko and Korenkov 1977) when certain Pseudomonas strains, isolated from chromate-containing sewage sludges, could reduce

Chromium–Microorganism Interactions

105

chromate, dichromate, and crocoite during anaerobic growth. Since then, several bacteria with exceptional ability to reduce Cr(VI) have been isolated from Crcontaminated and uncontaminated soil samples. Microorganisms implicated in direct or indirect reduction of Cr(VI) are listed in Table 1. Cr(VI) Reduction in Microbial Cultures Since the first reports of isolation of facultative anaerobic Cr(VI)-reducing bacteria in the mid-1970s (Romanenko and Korenkov 1977), the literature is abundant with instances of the reduction of Cr(VI) by several microorganisms, bacteria in particular (Ohtake and Silver 1995), mostly isolated from Cr-impacted environments (Table 1). Strains of Oscillatoria, Chlorella, and Zoogloea have also been reported to enzymatically reduce Cr(VI) (Losi et al. 1994b). But, as noticed with bacterial resistance to Cr(VI), Cr(VI)-reducing bacteria have been isolated also from environments with minimum or no impact of Cr (Wang et al. 1989; Turick et al. 1996; Badar et al. 2000). It is also interesting to note that pure cultures of microorganisms not previously exposed to Cr(VI) were capable of reducing it (Gvozdyak et al. 1986). Although the exact mechanism is not known, microorganisms capable of reducing Cr(VI) acquired the enzymes for degrading related compounds present in the environment or produce the reductants that in turn reduce Cr(VI) by chemical redox reactions. Anaerobic chromate-reducing strains are prevalent in subsurface soils and probably enhance Cr reduction in this environment (Turick et al. 1996). Chromium(VI) resistance and reduction are not necessarily interlinked. Chromium(VI) may be reduced by both Cr(VI)-resistant and Cr(VI)-sensitive strains of bacteria, and not necessarily all Cr-resistant bacteria can reduce Cr(VI). For instance, some aerobic Cr(VI)-resistant bacteria were not capable of reducing it (Gvozdyak et al. 1986; Wang et al. 1989). Cr(VI) reduction in aerobic conditions may not be a resistance mechanism in bacteria, but a trivial side activity of the reductase that may have evolved on other substrates (Ishibashi et al. 1990). In a bioprocess strategy for effective bioremediation of Cr(VI), it is important to use Cr(VI)-resistant microbes, with ability to reduce it. Two strains of Pseudomonas fluorescens, one resistant and the other sensitive to Cr(VI), reduced Cr(VI) at comparable rates (Ohtake et al. 1987; Bopp and Ehrlich 1988). Likewise, three Cr(VI)-sensitive bacteria from an uncontaminated soil and three Cr(VI)-resistant bacteria from two metal-stressed foundry soils and a tannery readily reduced Cr(VI) anaerobically (Badar et al. 2000). Interestingly, a Cr(VI)-sensitive Bacillus sp. from the uncontaminated soil was the most effective in reducing Cr(VI) among the three Cr(VI)-resistant bacterial strains from metal-stressed soils and three Cr(VI)-sensitive bacterial strains from the uncontaminated soil. These bacteria grew aerobically in acetate minimal medium supplemented with sodium chromate, but reduced Cr(VI) only anaerobically in the suspension of resting cells of aerobically grown bacteria. Anaerobic growth of the bacterium at the expense of Cr(VI) as electron acceptor was negligible. Conversely, an Arthrobacter sp., isolated from a long-term tannery wastecontaminated soil, was resistant to Cr(VI) at 100 µg/mL but could not reduce it

106

Table 1. Microorganisms capable of reducing Cr(VI).

Identification

Source Metal refining wastewaters

Enterobacter cloacae HO1

Activated sludge

Enterobacter cloacae HO1

Enterobacter cloacae HO1

2500 mg Cr(VI)/L

Cr-reducing efficiency

Mechanism

Reference

80%–95% from 50– 2000 ppm

Indirect; involving H2S Fude et al. 1994

520 mg/L

90% removal

Membrane associated, CrO2− 4 as terminal electron acceptor

Activated sludge

520 mg/L

90% removal

Anaerobic reduction of Komori et al. 1989; Cr(VI) while growOhtake et al. 1990; ing on acetate, maRege et al. 1997 late, succinate, ethanol, and glycerol, but inhibited by glucose

Activated sludge

520 mg/L

Wang et al. 1989; Komori et al. 1990a,b

Reduces Cr(VI) anaero- Wang et al. 1990, bically with acetate 1991; Fujii et al. and casamino acids 1990 as electron donors; O2 inhibits Cr(VI) reduction; activity associated with membrane fraction involving NADH as electron donor cytochrom c548

S.P.B. Kamaludeen et al.

Consortium of sulfatereducing bacteria (SRB)

Cr tolerance level

Table 1. (Continued). Identification

Source

Cr tolerance level

Cr-reducing efficiency

Aeromonas dechromatica

Bacillus sp. QC1–2

Soil and water samples near chromium-processing factory

Reference

Cr(VI) reduced in the Kvasnikov et al. 1985, presence of propyl 1986, 1987 alcohols and acetate; certain heavy metals promoted or inhibited Cr(VI) reduction/anaerobic Acetate, glucose/anaer- Gvozdyak et al. 1986 obic

—

17 mg/L

Sulfate starvation led to increased chromate uptake and chromate reduction 100% within 22 hr

Glucose/aerobic Peptone, glucose, H2/ anaerobic —

Wang and Xiao 1995 Romanenko and Korenkov 1977 Peitzsch et al. 1998

107

Cr(VI) reduction by Campos et al. 1995 cell-free extract; aerobically/anaerobically; Cr(VI) reductase (soluble NADH dependent)

Chromium–Microorganism Interactions

Achromobacter eurydice, Bacillus cereus, B. subtilis, Micrococcus roseus, Pseudomonas aeruginosa Bacillus sp. Pseudomonas dechromaticans Alcaligenes eutrophus — AE104

Mechanism

108

Table 1. (Continued). Identification

Source

Bacillus sp.

Chromatecontaminated soil

Agrobacterium radiobacter EPS-916

Soil

10–200 mg/L

26 mg/L

1.9 mg/L

Cr-reducing efficiency

Mechanism

Reference

Continuous-flow bioChirwa and Wang film reactor, 100% 1997a; Wang and removal in 6–24 hr Shen 1997 100% removal within Cr(VI) reduction under Llovera et al. 1993 6 hr both aerobic and anaerobic conditions 100% reduction — Krauter et al. 1996

—

—

Glucose/aerobic

—

—

—

Pseudomonas ambigua G-1

—

—

—

Pseudomonas stutzeri (two strains)

Metal-contaminated foundry soil

Several organic compounds as electron donors/anaerobic Aerobically reduced Shimada and MatsuCr(VI) while growshima 1983 ing on glucose Nutrient broth/aerobic; Horitsu et al. 1987; SuCr(VI)-reductase zuki et al. 1992 NAD(P)H-dependent Enzymatic NAD(P)H- Badar et al. 2000 dependent, Crresistant

Pseudomonas chromatophila Pseudomonas K-21

52 mg chromate/L

58%–88% Cr reduction anaerobically

Bopp and Ehrlich 1988; Chirwa and Wang 1997b; Wang and Shen 1997; Deleo and Ehrlich 1994 Lebedeva and Lyalikova 1979

S.P.B. Kamaludeen et al.

Saccharomyces cerevis- — iae Pseudomonas fluorescens — LB300

Cr tolerance level

Table 1. (Continued). Cr tolerance level

Cr-reducing efficiency

Tannery

52 mg/L

Uncontaminated soil

<5 mg/L

76% Cr reduction anaerobically 75%–92% reduction anaerobically

Identification

—

—

—

—

Pseudomonas aeruginosa Tanning effluent A2Chr

Pseudomonas sp.

Reference

Cr-resistant, grew aero- Badar et al. 2000 bically Cr-sensitive, grew aero- Badar et al. 2000 bically

Cr(VI) reduction, aero- Whole cells, cell susbically pension, cell-free extract — Cr(VI) reductase gene transferred to tobacco plant cells using Agrobacterium tumefaciens binary vector system Cr(VI) reduction, aero- In batch culture, dialbically ysis reactor, and cells entrapped in a biofilm in a biological rotating contactor; maximum reduction at 10 mg/mL Cr(IV) reduction, aero- In continuous susbically pended-growth cultures Cr(VI) reduction Cr(VI) reduction Acetate/anaerobic

Ishibashi et al. 1990 Jin et al. 2001

Khare et al. 1997; Ganguli and Tripathy 1999, 2001, 2002

Gopalan and Veeramani 1994 Bhide et al. 1996 Das and Chandra 1990 Kvasnikov et al. 1988

109

Pseudomonas mendocina Cooling tower effluent Streptomyces sp. — Escherichia coli

10–25 mg/mL

Mechanism

Chromium–Microorganism Interactions

An unidentified bacterium Pseudomonas synxantha, Bacillus sp. and an unidentified gram-positive strain Pseudomonas putida, P. fluorescens, and E. coli Pseudomonas aeruginosa HP014

Source

Identification Escherichia coli ATCC 33456

110

Table 1. (Continued). Source Activated sludge

Desulfovibrio vulgaris — ATCC29579, D. sulfuricans Pantoea agglomerans Surface sediments SP 1 Pantoea agglomerans SP 1

Surface sediments

Cr-reducing efficiency

16 mg/L

100% reduction in 12 hr

Mechanism

Reference

Enzymatic reduction/ Shen and Wang 1993; glucose, acetate, proWang and Shen pionate/aerobic and 1997 anaerobic A wide range of elec- Shen and Wang 1993, tron donors for 1995; Wang and Cr(VI) reduction Xiao 1995

Continuous-flow biore- Wang and Chirwa actor 1998

—

—

—

—

—

—

—

—

Reduces Cr(VI), Fe(III), U(VI), and Te(VII) c3 cytochrome as Cr(VI) reductase/H2/ anaerobic Dissimilatory reduction on acetate and other electron donors Elemental sulfur disproportionate dissimilatory reduction under anaerobic conditions

Fredrickson et al. 2000 Lovley 1993; Lovley and Phillips 1994 Francis et al. 2000 Obraztsova et al. 2002

S.P.B. Kamaludeen et al.

Coculture of Escherichia Both from activated coli ATCC 33456 sludge [Cr(VI) reducer] and Pseudomonas putida DMP-1 (phenol degrader) Coculture of Escherichia Both from activated coli ATCC 33456 sludge [Cr(VI) reducer] and Pseudomonas putida DMP-1 (phenol degrader) Deinococcus radiodurans —

Cr tolerance level

Table 1. (Continued). Identification

Source

Cr tolerance level

Cr-reducing efficiency −1

Anaerobic enrichment

Aquifer sediments

Desulfotomaculum reducens

Cr-contaminated sediments

39 mg/L

73%–94% reduction anaerobically and 18%–63% reduction aerobically

Reference

Both anaerobically and Srinath et al. 2001 aerobically in peptone water

Shakoori et al. 1999, 2000 McLean and Beveridge 2001; McLean et al. 2000

Marsh and McInerney 2001 Tebo and Obraztsova 1998

111

87% of the Cr(VI) in — 20 mg K2Cr2O7/mL reduced in 72 hr 23% reduced in 24 hr Respire aerobically and 19% in the presence anaerobically using a of As(V), 27% in variety of terminal the presence of electron acceptors, Cu(II) Fe(III), Mn(IV), NO−2, NO−3, SO2; soluble enzyme, mainly cytoplasmic, involved in reduction 39 mg/L in 6 d Cr(VI) reduction and growth were dependent on H2 10 mg/L Grows and reduces Cr(VI) as sole electron acceptor in the presence of butyrate as the carbon source and in the absence of sulfate; no growth in absence of Cr(VI)

Chromium–Microorganism Interactions

Eight facultative anaerTannery effluents >400 µg chromate mL obes (five isolates of Aerococcus, two isolates of Micrococcus, and one isolate of Aeromonas) Gram-positive dichroTannery effluent 80 mg K2Cr2O7/mL mate-resistant bacterium (ATCC 700729) Pseudomonas sp. CRB5 Wood preservation 520 mg Cr(VI)/L site with chromated copper arsenate

Mechanism

112

Table 1. (Continued).

Identification

Source

Cr tolerance level

Cr-reducing efficiency

Mechanism

Reference

Anaerobic zone of Mn-rich sediments

—

Cr(V) detected during Reduces a diverse Myers et al. 2000 reduction by forarray of compounds mate-dependent under anaerobic conCr(VI) reductase ditions including Mn(IV), Fe(III), fumarate, and Cr(VI); formate- or NADHdependent Cr(VI) reductase activity in anaerobically grown cells with highest activity in cytoplasmic membrane

Shewanella oneidensis

Anaerobic zone of Mn-rich sediments

—

Live, resting cells re- Precipitates encrusting duced 80% of bacterial cells conCr(VI) in 1 hr; sotained Cr(III) or dium azide and heat lower oxidation treatment stopped states reduction; no reduction in cell-free supernatants

Shewanella oneidensis

Anaerobic zone of Mn-rich sediments

—

Cr(VI) reduction under fumarate- and nitrate-reducing conditions

Daulton et al. 2002

Cr(VI) inducible, asso- Viamajala et al. ciated with anaerobic 2002a,b growth of bacterium

S.P.B. Kamaludeen et al.

Shewanella oneidensis (earlier designated as S. putrifaciens MR-1)

Table 1. (Continued).

Identification

Source

Thiobacillus ferroxidans

—

Cr-reducing efficiency

Mechanism

Reference

Cr(VI) reduction by Sisti et al. 1996, 1998 protons (sulfite, thiosulfate, etc.) with high reducing power generated during oxidation of elemental sulfur; aerobic

—

Cr(VI) reduction by Sisti et al. 1996, 1998; protons (sulfite, thioQuiIntana et al. sulfate, etc.) with 2001 high reducing power generated during oxidation of elemental sulfur; related to the concentration of protons generated; occurs at pH 2–8, more pronounced at lower pH; aerobic and anaerobic, but more pronounced aerobically

113

—

Chromium–Microorganism Interactions

Thiobacillus ferroxidans

Cr tolerance level

114

Table 1. (Continued). Identification Bacillus subtilis Sporosarcina ureae Shewanella putrefaciens

Source

Cr-reducing efficiency —

—

—

—

—

Chlorella

—

—

Anabaena variabilis

—

—

Transitory formation of Cr(V) during Cr(VI) reduction Enzymatically reduced Cr(VI) Enzymatically reduced Cr(VI) Chromate reduced in 18 d to Cr(III); 50% of Cr(III) formed accumulated in the cells and 50% in the medium; Cr(VI) reduction associated with heterocysts

Mechanism

Reference

Nonmetabolizing bacte- Fein et al. 2002 rial cells reduced Cr(VI) on cell surface in absence of externally supplied electron donors; not coupled to oxidation of bacterial exudates; fastest under acidic conditions — Liu et al. 1995

—

Losi et al. 1994b

—

Losi et al. 1994b Garnham and Green 1995

S.P.B. Kamaludeen et al.

—

Algae Spirogyra sp. and Mougeotia sp. Oscillatoria

—

Cr tolerance level

Chromium–Microorganism Interactions

115

at this concentration (Megharaj et al. 2003). Likewise, Cr(VI) reduction occurred equally rapidly with both Cr(VI)-resistant and plasmid-cured Cr(VI)-sensitive strains of P. fluorescens (Bopp and Ehrlich 1988). Chromate resistance determinants have been described on plasmids in several bacteria, especially in Pseudomonas. However, Cr(VI) reduction determinants have not been found on plasmids. Cr(VI) reduction was independent of chromate resistance, conferred by plasmid pLHB1, in P. fluorescens (Bopp and Ehrlich 1988). In P. mendocina (Bhide et al. 1996), however, plasmid pAR1180 determined both chromate resistance and Cr(VI) reduction (Dhakephalkar et al. 1996). Microorganisms that are known to reduce Cr(VI) reduce it under aerobic or anaerobic conditions, but the physiological role in such transformations is not clear. Earlier reports (Romanenko and Korenkov 1977; Lebedeva and Lyalikova 1979; Kvasnikov et al. 1985) have shown that facultative anaerobes (Pseudomonas and Aeromonas strains) reduce Cr(VI) to Cr(III) anaerobically. Anaerobic bacteria with great Cr(VI)-reducing potential are ubiquitous in both Cr(VI)contaminated and uncontaminated soils (Turick et al. 1996; Schmieman et al. 1998). There is no convincing evidence yet to suggest that Cr(VI) serves as the electron acceptor to support the anaerobic growth of bacteria. Enterobacter cloacae grew well under aerobic conditions and slowly under anaerobic conditions at chromate concentrations above 10 mM in nutrient broth, but could reduce chromate only under anaerobic conditions (Wang et al. 1989). Also, there is evidence that O2 inhibited the reduction of Cr(VI) by E. cloacae strain HO1 in a medium containing other carbon sources as electron donors (Wang et al. 1989; Komori et al. 1990a, b). Likewise, Escherichia coli could reduce Cr(VI) only in the absence of O2 (Shen and Wang 1994a,b). Under anaerobic conditions, Cr(VI) serves as a terminal electron acceptor through electron transport systems involving cytochrome c in Enterobacter cloacae (Wang et al. 1989), cytochrome b and d in Escherichia coli (Shen and Wang 1993) and cytochrome c3 in Desulfovibrio vulgaris (Lovley and Phillips 1994). Membrane or soluble fractions may be involved in the reduction of Cr(VI). Under aerobic conditions, both NADH and endogenous cell reserves may serve as elecron donors for Cr(VI) reduction. A recent study (Marsh and McInerney 2001) established a relationship between the bioavailability of H2 and chromate reduction in anaerobic aquifer sediments. The anaerobic enrichment, developed from the sediment, utilized Cr(VI) and was dependent on H2 for growth and chromate reduction. In the absence of Cr(VI), H2 accumulated in the anaerobic medium. Under Cr(VI)-reducing conditions, however, no H2 and methane accumulated because H2 was utilized by the enrichment. When H2 was provided in the medium as the electron donor, the enrichment could reduce 40 mg/L Cr(VI) in 6 d. Increasing the availability of H2 by addition of suitable electron donors (formate, H2 and glucose) accelerated the reduction of chromate. Gram-positive bacteria, capable of reducing Cr(VI) as a terminal electron acceptor and with a relatively high level of resistance to chromate, have been isolated from tannery effluents (Basu et al. 1997; Shakoori et al. 1999, 2000). A chromate-resistant gram-positive bacterium (ATCC 700729) tolerated high

116

S.P.B. Kamaludeen et al.

concentrations (up to 80 mg/mL) of dichromate and reduced 87% of the Cr(VI) in 20 mg K2Cr2O7/mL in 72 hr in a nutrient-rich medium (Shakoori et al. 2000). The bacterium could reduce Cr(VI) even at a concentration of dichromate as high as 80 mg/mL, but the reduction required a longer time at 80 mg/mL than at 20 mg/mL. Chromate reduction occurs either anaerobically (Bopp and Ehrlich 1988; Wang et al. 1989; Wang and Shen 1997; Badar et al. 2000; Srinath et al. 2001), aerobically (Ishibashi et al. 1990; Wang and Shen 1997), and under both conditions (McLean and Beveridge 2001) (see Table 1). Agrobacterium radiobacter EPS-916 (Llovera et al. 1993) and Escherichia coli ATCC 33456 could reduce Cr(VI) under both aerobic and anaerobic conditions. Likewise, a pseudomonad, isolated from a wood preservation site contaminated with chromated copper arsenate, reduced chromate under both aerobic and anaerobic conditions (McLean and Beveridge 2001). Srinath et al. (2001) also reported that Cr(VI)-tolerant (>400 µg/mL) facultative anaerobes (five isolates of Aerococcus sp., two isolates of Micrococcus sp., and one isolate of Aeromonas sp.), isolated from tannery effluent, apparently reduced Cr(VI) both anaerobically and aerobically. Cr(VI) reduction by these facultative anaerobes in diluted peptone water was more pronounced under anaerobic conditions (73%–94% reduction) than under aerobic conditions (18%– 63% reduction), but conditions used for anaerobic and aerobic systems have not been described. Because peptone alone may catalyze chemical reduction of Cr(VI) (Wang and Shen 1995), it was not clear whether the Cr(VI) reduction in microbial cultures was caused chemically, microbially, or both. Cell suspensions of Pseudomonas putida PRS 2000, P. fluorescens LB303, and Escherichia coli AC80 aerobically reduced Cr(VI) to Cr(III) (Ishibashi et al. 1990). Reduction of Cr(VI) in cell suspensions of these bacteria was more rapid and complete aerobically than anaerobically. After disruption of the cells of P. putida and centrifugation, the supernatant, but not the membrane fraction (pellet), reduced all the added Cr(VI) within 1 hr. Likewise, Wang and Shen (1997) reported that resting cells of Bacillus sp. and Pseudomonas fluorescens LB300 aerobically reduced Cr(VI). However, Cr(VI) reduction by the cells of Escherichia coli was inhibited in the presence of oxygen (Shen and Wang 1994a,b). Enterobacter cloacae, a chromate-resistant strain, could grow in the presence of Cr(VI) under both aerobic and anaerobic conditions, but Cr(VI) was reduced only anaerobically (Wang et al. 1989). The strain lost both resistance and Cr(VI)-reducing ability in anaerobic growth on nitrate. Cifuentes et al. (1996) reported that organic amendments enhanced the reduction of Cr in soils by indigenous microflora. Generally, Cr(VI) reduction by growing bacterial cells has been demonstrated in media containing natural aliphatic compounds, amino acids, and fatty acids as electron donors (Wang and Shen 1995). Microbial reduction of Cr(VI) occurred during anaerobic degradation of benzoate (Shen et al. 1996). A dissimilatory metal-reducing bacterium, Shewanella oneidensis, could reduce Cr(VI) when grown on fumarate or nitrate as an electron acceptor and lactate as an electron donor (Viamajala et al. 2002a). Cr(VI) reduction under fumarate- and denitrifying conditions, dependent on the

Chromium–Microorganism Interactions

117

physiological state of the culture, was possibly inducible under anaerobic conditions. Cr(VI) reduction in the anaerobically grown stationary phase of this bacterium is a complex process, possibly involving more than one pathway (Viamajala et al. 2002b). A wide range of organic pollutants such as phenol, 2-chlorophenol, p-cresol, 2,6-dimethylphenol, 3,5-dimethylphenol, 3,4-dimethylphenol, benzene, and toluene can also serve as electron donors for Cr(VI) reduction in cocultures containing E. coli ATCC33456 and P. putida DMP-1 (Shen and Wang 1995). Metabolites produced during phenol degradation by P. putida served as electron donors for Cr(VI) reduction by E. coli. Technology using such cocultures would help to simultaneously detoxify both organic pollutants and the toxic Cr(VI). Nonmetabolizing resting cells of bacteria could reduce Cr(VI), but only in the presence of an added carbon source (Bopp and Ehrlich 1988; Shen and Wang 1994b; Philip et al. 1998). Killed resting cells could not cause Cr(VI) reduction (Shen and Wang 1994b; Wang and Shen 1997). Soluble enzymes in cell extracts can reduce Cr(VI) in the presence (Horitsu et al. 1987; Philip et al. 1998) or absence (Bopp and Ehrlich 1988; Shen and Wang 1994b) of added electron donors. According to very recent evidence, nonmetabolic Cr(VI) reduction can occur on bacterial surfaces even in the absence of externally added electron donors in the medium. Thus, Fein et al. (2002) demonstrated that nonmetabolizing cells of Bacillus subtilis, Sporosarcina ureae, and Shewanella putrefaciens could reduce significant amounts of Cr(VI) in the absence of externally supplied electron donors. The Cr(VI) reduction by the bacterial strains was dependent on solution pH, decreasing with increasing pH, and presumably occurred at the cell wall and independent of the oxidation of bacterial organic exudates. Such nonmetabolizing reduction of Cr(VI) by bacteria in nutrient-poor conditions may be important in the biogeochemical distribution of Cr. Cr(VI) reduction by microorganisms, known to occur under both aerobic and anaerobic conditions (see Table 1), is a redox-sensitive process (Shen and Wang 1994b; Chen and Hao 1996). The ability of washed resting cells of Agrobacterium radiobacter EPS-916 to reduce Cr(VI) was governed by their redox potenial (Llovera et al. 1993). Resting cells of A. radiobacter EPS-916, pregrown under aerobic conditions on glucose, fructose, maltose, lactose, mannitol, or glycerol as the sole carbon and energy source, exhibited similar redox potentials of around −200 mV and completely reduced 0.5 mM chromate. On the other hand, the inability of the resting cells of the bacterium, pregrown on glutamate or succinate, to reduce chromate was associated with relatively high redox potentials of −138 to −132 mV. Moreover, resting cells pregrown under anaerobic conditions on glucose had lower redox potentials (−240 mV) and a more pronounced chromate-reducing activity than did the aerobically grown resting cells on glucose with a redox potential of −200 mV. Likewise, cells pregrown anaerobically on chromate as the electron acceptor effected more rapid reduction of chromate than did the anaeorobically grown cells (−198 mV) on nitrate. Evidence suggested a negative correlation between chromate reduction by the rest-

118

S.P.B. Kamaludeen et al.

ing cells of A. radiobacter EPS-916 and their redox potential. On the other hand, in an anaerobic enrichment from aquifer sediment, Cr(VI) reduction appears to occur under nitrate-reducing conditions but before iron and sulfate reduction (Marsh and McInerney 2001). Evidently, highly reducing conditions, necessary for the reduction of iron and sulfate and methanogenesis, may not be required for chromate reduction. Abiotic reduction of Cr(VI) has also been demostrated in media rich in nutrients containing some reductants, especially under predominantly reduced conditions. Thus, even in sterile tryptic soy broth, Cr(VI) was reduced abiotically with time as a function of redox potential (Turick et al. 1996). Thus, more than 50% of 25 µg Cr(VI)/mL added to the tryptic sterile broth was reduced abiotically in 60–80 hr at redox potentials of −120 and −380 mV, as compared to <27% reduction at +243, +186, and +58 mV during the corresponding period. It is therefore necessary to have appropriate control to exclude the chemical redox reactions when nutrient-rich growth media are used to assess the Cr(VI)reducing ability of pure cultures of microorganisms. Cr(VI) Reductases Cr(VI) reduction is mediated enzymatically (direct) or nonenzymatically (indirect). There is considerable literature on the involvement of Cr(VI) reductases in direct reduction of Cr(VI) to Cr(III) by bacteria. In growing cultures with added carbon sources as electron donors and in cell suspensions, Cr(VI) reduction can be predominantly aerobic or anaerobic, but generally not both. Interestingly, Cr(VI) reductases can catalyse reduction of Cr(VI) to Cr(III) anaerobically (Lovley and Philipps 1994), aerobically (Ishibashi et al. 1990; Suzuki et al. 1992), and also both anaerobically and aerobically (Bopp and Ehrlich 1988; Shen and Wang 1993; Campos-Garcia et al. 1997; McLean and Beveridge 2001). The Cr(VI) reductase enzyme may be present in the membrane fraction of the cells as in Pseudomonas fluorescens and Enterobacter cloacae (Wang et al. 1990) or in the soluble fraction of the cells (cell-free system) as in P. ambigua (Horitsu et al. 1987), P. putida (Ishibashi et al. 1990), and a Bacillus sp. (Campos et al. 1995), with NADH, NADPH or H2 (Desulfovibrio vulgaris) as electron donors and possible involvement of cytochromes b, c, and d. Membrane vesicles of E. cloacae, reduced with NADH and then exposed to Cr(VI), oxidized cytochromes c and b and reduced Cr(VI). Evidence suggested that cytochrome c548 specifically was involved in the reduction of Cr(VI) by membrane vesicles (Wang et al. 1991). In the presence of H2 and excess of hydrogenase, cytochrome c3, a periplasmic protein, in the soluble cell-free fraction of the cells in D. vulgaris (Lovley and Phillips 1994) reduced Cr(VI) 50 times faster than did the Cr(VI) reductase of P. ambigua with NADH and NADPH as electron donor (Horitsu et al. 1987). Soluble fractions of the cell-free extract, largely cytoplasmic, of a pseudomonad from a wood preservation site reduced chromate, added at 10 mg Cr(VI)/L, under both aerobic (55%) and anaerobic (80%) conditions in 2.5 hr (McLean and Beveridge 2001). Cr(VI) reductase in anaerobically grown Shewanella putrefaciens MR-1 was formate dependent with highest activity in the cytoplasmic membrane (Myers et al. 2000). The Cr(VI) reduc-

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tase in P. ambigua (Suzuki et al. 1992) and a Bacillus sp. (Campos-Garcia et al. 1997) have been purified and characterized. More recently, to clone a chromate reductase gene, a novel soluble chromate reductase of P. putida has been first purified to homogeneity and characterized, using ammonium sulfate precipitation, anion-exchange chromatography, chromatofusing, and gel filtration (Park et al. 2000). The reductase activity was NADH- or NADPH dependent. The optimum conditions for the chromate reductase were 80 °C and pH 5.0. Kinetic properties of the enzyme showed Km of 374 µM CrO42− and Vmax of 1.72 µmol/ min/g protein. Suzuki et al. (as cited in Park et al. 2000) sequenced the gene encoding the chromate reductase (Suzuki et al. 1992) from P. ambigua. The genes encoding the chromate reductase in P. ambigua and P. putida were not homologous, however. Reduction Products Generally, in bacterial cultures or in enzyme systems, Cr(VI) is reduced to Cr(III) without transitory accumulation of any intermediate, but there are instances when Cr(V) accumulates as a transitory intermediate during microbial conversion of Cr(VI) to Cr(III). For instance, the NADPHdependent Cr(VI) reductase in P. ambigua catalyzed the transitory formation of Cr(V) during conversion of Cr(VI) to Cr(III) (Suzuki et al. 1992). Toxicity of Cr(VI) to microorganisms is probably associated with the transient formation of Cr(V) as an intermediate. Cr(V) is formed also during reduction of Cr(VI) in algal cultures and in reactions with physiological reducing agents such as NADPH, glutathione, and several pentoses (Shi and Dalal 1990). In most studies, conclusions on microbial reduction of Cr(VI) were based on its disappearance or accumulation of the Cr(III) [determined as the difference in total Cr and Cr(VI)] as the reduction product with incubation. The colorimetric diphenylcarbazide method commonly used in Cr(VI) estimation is not specific because its probable reduction product Cr(V) and hexavalent forms of Mo, V, and Hg can also react with the same reagent. The direct measurement of the oxidation state of the Cr during bacterial reduction of Cr(VI), however, has not been attempted until recently. Daulton et al. (2002) used the electron energy loss spectroscopy (EELS) technique to characterize the oxidation state of Cr during Cr(VI) reduction by Shewanella oneidensis in anaerobic cultures. Transmission electron microscopy (TEM) of the cells exposed to Cr(VI) showed that the cells were encrusted in Cr-rich precipitates, mostly restricted to the outer surface of the cells. These precipitates, based on analysis by EELS, contained Cr(III) or its lower state of oxidation. Myers et al. (2000), using electron paramagnetic resonance (EPR) spectroscopy, confirmed the formation of Cr(V) via one-electron reduction of Cr(VI) as the first step by a facultative anaerobe, Shewanella putrefaciens MR-1. C. Indirect Reduction Apart from the direct (enzymatic) reduction of Cr(VI) as discussed under III.B, microorganisms can also mediate the reduction of Cr(VI) indirectly, involving a biotic–abiotic coupling. For instance, Fe(II) and S2−, produced by microorgan-

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isms through dissimilatory reduction pathways, can chemically catalyze several biogeochemical processes including Cr(VI) reduction (Lovley 1993; Fendorf and Li 1996). Fe(III), an important electron acceptor for microbial oxidation of organic compounds (aliphatic and aromatic), is one of the most abundant metals in the soil. A wide range of bacteria couple the oxidation of organic compounds and H2 to reduction of Fe(III) and SO4 to Fe(II) and H2S, respectively, under oxygen stress conditions (Lovley 1993); this reaction occurs in submerged rice soils, for example. Different genera of Fe(III)-reducing bacteria reduce Fe(III) via different mechanisms (Nevin and Lovley 2002). Recently, Wielinga et al. (2001) demonstrated the reduction of Cr(VI) by a biotic–abiotic coupling mechanism involing iron reduction. Dissimilar Fe(III) reduction by Shewanella alga ATCC 51181, a facultative anaerobic bacterium, under iron-reducing conditions provided a primary pathway for chemical reduction of Cr(VI), injected into a bioreactor, by microbially induced ferrous ion. However, it has been difficult to differentiate the exact contribution between biological (direct) and chemical (indirect: biotic–abiotic) reduction of Cr(VI) in a soil environment. Evidence using Desulfovibrio vulgaris as a model chromate reducer suggests that chemical reduction of chromate by Fe(II) was 100 times faster than that by D. vulgaris, a chromate reducer (Wielinga et al. 2001). In anaerobic environments abundant in Fe(II), nonenzymatic reduction of Cr(VI) by Fe(II) can be as important as enzymatic Cr(VI) reduction (Masscheleyn et al. 1992). A facultative anaerobe, Pantoea agglomerans SP1, coupled anaerobic growth on acetate and other electron donors to the dissimilatory reduction of electron acceptors, Fe(III), Mn(IV), and Cr(VI), but not sulfate (Francis et al. 2000). When Cr(VI) was added to this γ-protobacterium culture with elemental sulfur alone, S0-disproportionation to sulfate and hydrogen sulfide occurred with concomitant growth of the bacterium and reduction of Cr(VI) (Obraztsova et al. 2002). Likewise, P. agglomerans SP1 grew chemolithoautotrophically by the S0-disproportionation, coupled to reduction of Fe(III) and Mn(IV). S0-Disproportionation that may be widespread in certain anaerobic environments may provide an effective mechanism for attenuation of Cr(VI) through its reductive detoxification. Sulfate-reducing bacteria (obligate anaerobic heterotrophs) couple the oxidation of organic sources to the reduction of sulfate to sulfide. Reduction of Cr(VI) by bacterially produced hydrogen sulfide, followed by precipitation of the Cr(III) formed, is an important mechanism in sulfate-rich soil environments when anaerobic conditions prevail (Losi et al. 1994c; Pettine et al. 1994, 1998), as in flooded compacted soils. Likewise, sulfide produced by sulfate-reducing bacteria has been implicated in Cr(VI) reduction in marine environments (Smillie et al. 1981). Hydrogen sulfide produced in acid sulfate soil under reducing conditions is easily precipitated as FeS in reduced soils (Ponnamperuma 1972) and sediments. Fe(II) (Eary and Rai 1988) and H2S (Pettine et al. 1994), both microbially produced, are effective reductants of Cr(VI) under reduced conditions as is the FeS (Patterson et al. 1997). Low concentrations of Cr(VI) can accelerate the growth and sulfate-reducing activity of sulfate-reducing bacteria (Karnachuk

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1995) and thereby the reduction of Cr(VI) by the H2S evolved. A spore-forming sulfate-reducing bacterium, Desulfotomaculum reducens sp. nov. strain MI-1, isolated from sediments with high concentrations of Cr and other heavy metals by enrichment, could grow with Cr(VI) as sole electron acceptor in the absence of sulfate with butyrate, lactate, or valerate as the electron donor (Tebo and Obraztsova 1998). Cr(VI) was presumably reduced to Cr(III) as Cr(OH)3. In the absence of Cr(VI), no bacterial growth was noticed. Biologically generated sulfur compounds with high reducing power such as sulfite, thiosulfate, and polythionate can catalyze the chemical reduction of Cr(VI). Chemoautotrophic bacteria belonging to the Thiobacilli group, which can derive energy from the oxidation of inorganic sulfur compounds during sulfur oxidation, generate a range of sulfur compounds such as sulfite and thiosulfate with high reducing power that can in turn catalyze the reduction of Cr(VI). For instance, Thiobacillus ferroxidans grown on elemental sulfur has been used to reduce Cr(VI) under aerobic conditions (Sisti et al. 1996, 1998). The Cr(VI)reducing ability of the cells of this bacterium under aerobic conditions in shake flasks and in fermentation vessels was related to the generation of protons with high reducing power from elemental sulfur (QuiIntana et al. 2001). T. ferroxidans could reduce Cr(VI) over a wide pH range (2–8), interestingly with more pronounced reduction at lower pH, associated with increased oxidation of elemental sulfur to products with high reducing power. Cr(VI) reduction, mediated by T. ferroxidans in the presence of elemental sulfur, occurred under both aerobic and anaerobic conditions, but more effectively under aerobic conditions. Evidently bacterial reduction of Cr(VI), involving biotic–abiotic coupling, can occur under both sulfate-reducing and sulfur-oxidizing conditions. Thus, Cr(VI) reduction or immobilization can be effected abiotically by different substances, but there has been considerable progress in recent years on the feasibility of using biological reduction for treatment of Cr(VI)-containing wastes. D. Biotic–Abiotic Coupling in Mn Oxide-Mediated Oxidation of Cr(III) Although an extensive literature exists on microbial reduction of Cr(VI), direct oxidation of Cr(III) to Cr(VI) in pure cultures of microorganisms has not been demonstrated yet. Because Cr(VI) is toxic to most microorganisms because of its high oxidizing nature, direct oxidation of Cr(III) by microorganisms may not be a predominant process. However, indirect involvement of microorganisms in the oxidation of Cr(III) to more toxic Cr(VI) through biotic–abiotic coupling is of great concern in environments rich in easily reducible forms of Mn. In indirect involvement of microorganisms in Cr oxidation, microorganisms first oxidize Mn(II) to higher states of oxidation, in particular that, in turn, chemically oxidize the Cr(III) generally to Cr(VI). Oxidation of Cr(III) in the soil involving biotic–abiotic coupling could involve essentially Mn-reducing microorganisms that may indirectly facilitate chemical Cr oxidation (see Fig. 2). A diverse range of Fe(III) and Mn(IV) oxide forms is potentially formed by microorganisms to utilize energy during redox reactions and also to prevent the

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accumulation of Fe and Mn at toxic concentrations in the environment. Fe(III), available in significant quantities in the soil, is used as an electron acceptor in the degradation of organic pollutants (Lovley and Phillips 1988) and reduction of Cr(VI), as well as in anaerobic environments. Although it is speculated that both Fe and Mn redox systems could be involved in Cr oxidation reactions, the main oxidants of Cr(III) in soils are Mn oxides (Bartlett and James 1988). Fe-redox systems are probably involved more in the reduction of Cr(VI) to Cr(III) than in its oxidation. This section focuses on Mn systems, as the role of Mn oxides in chemical oxidation of Cr in soils is well established [see Avudainayagam et al., this volume (Section II.E)]. The nature and amount of Mn oxides formed in the soils determine the rate of Cr oxidation. However, oxidation of Mn(II) to Mn(IV) in the soil is mediated essentially by microorganisms. The following section discusses the processes of microbial Mn oxidation in soils, mechanisms of Mn oxidation, and the nature of Mn oxides formed, given that such oxides could play a significant role in Cr(III) oxidation, especially in long-term Cr-contaminated soils abundant in Mn. Microbial Oxidation of Manganese Microorganisms are responsible for much of the Mn(II) oxidation observed in the environment (Bromfield and Skerman 1950), contributing up to five orders of magnitude compared to abiotic Mn(II) oxidation (Wehrli et al. 1995). There is some evidence for oxidation of Mn(II) by chemical means (Ross and Bartlett 1981). The oxidation of Mn(II) to its higher oxides in natural environments such as soils (Leeper and Swaby 1940), freshwater lakes (Tipping et al. 1985), and estuarine waters and sediments (Edenborn et al. 1985) is mediated by a wide range of heterotrophic microorganisms (Ghiorse 1988). The microorganisms involved are capable of oxidizing Mn(II) far more effectively at neutral pH levels (pH 6–8) than any nonbiological system. Oxidation of Mn(II) to Mn(IV) is an exergonic reaction, yielding approximately −18.2 kcal mol−1 at 1 M concentrations of the reactants, which can be used as an energy source by microorganisms, but also there are many contradictory reports (Ghiorse 1984a,b). That this reaction minimizes the toxic effect of Mn(II) by transforming it to insoluble oxides, which cannot enter the cell, is yet another advantage (Bromfield 1976). The heterotrophic microbial populations that can effect the conversion of soluble Mn to solid Mn oxides include prosthecate bacteria, sheathed bacteria, fungi, algae, and their synergistic combinations (Ehrlich 1981; Nealson 1983; Nealson et al. 1988). Mn oxidation mechanisms may be either direct or indirect. The direct mechanisms involve either catalysis or specific binding of cell-associated materials, which enhance autoxidation later. Indirect mechanisms relate to microbially promoted changes in the cell microenvironment that later lead to nonbiological oxidation of Mn(II) (Greene and Madgwick 1991). Microbial Mn Oxidizers A wide range of microorganisms exhibit Mn-oxidizing ability in water and soil matrices: Bacillus SG1 (Rosson and Nealson 1982; Mandernack et al. 1995a), Arthrobacter sp. and Leptothrix sp. (Ghiorse 1984b),

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Pseudomonas sp. S36 (Nealson 1983), Oceanospirillum sp. (Ehrlich 1982), Pedomicrobium sp. (Larsen et al. 1999), the white rot fungus Phanerochaete chrysosporium (Kirk and Farrell 1987), and an isolate of Streptomyces (Bromfield 1979). Algae such as Scenedesmus (Knauer et al. 1999) and Chlamydomonas (Greene and Madgwick 1991; Stuetz et al. 1996) are also known to oxidize Mn. Mechanism of Mn Oxidation As already discussed, the oxidation of Mn is mediated either enzymatically or by its binding to other extracellular secretions (Table 2). The majority of the Mn-oxidizing systems are extracellular as in cultures of fungi and a Streptomyces, wherein the oxidizing factors are diffused into the surrounding environment (Bromfield 1979). The Mn oxides formed in fungus are accumulated mainly in the hyphae, resulting in black-colored colonies. The take-all fungus (Gaeumannomyces graminis var. tritici) has been reported to oxidize Mn in soils. Bacillus sp. and Leptothrix discophora have been extensively studied for their mode of Mn oxidation. In Bacillus SG1, the spore coats bind and oxidize Mn(II) in a manner similar to whole spores (deVrind et al. 1986). Greene and Madgwick (1991) have reported that a Pseudomonas strain in association with an alga (Chlamydomonas) oxidized 5 g Mn(II) L−1 to disordered semipure γ-MnO2 and manganite (γ-MnOOH) as an intermediate. Indirect Mn oxidation, where the changes in pH or Eh of an environment are modified as a result of metabolism and growth of microorganisms, is well documented. Bacteria and fungi have long been recognised to catalyze this type of nonspecific manganese oxidation (Ehrlich 1976). Microbial Mn Oxides Although microbial Mn(II) oxidation has been investigated extensively, less attention has been given to the characterization of the Table 2. Microorganisms capable of oxidising manganese.

Organism Bacillus sp. SG1 Leptothrix discophora Pseudomonas sp. S-36 Pedomicrobium sp. Scenedesmus sp.

Source

Oxidizing part

Shore sediments Mature dormant spors Freshwater sedi- Protein in ments sheaths (110 kDa) Marine sediExtracellular ment enrichglycocalyx ment Freshwater Extracellular enzymes Freshwater Extracellular

Optimum temperature for oxidation

pH

Reference

4 °–45 °C

7.8

Rosson and Nealson 1982 Ghiorse 1984b

7.8 °–28 °C

Nealson 1978

25 °C

7.9

Larsen et al. 1999 Knauer et al. 1999

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microbially formed Mn oxides (Greene and Madgwick 1991). Generally, the identification of Mn oxides is problematic because of their complex nature. The most common methods used for characterization of Mn oxides are X-ray diffraction and Fourier transform infrared (FTIR) spectroscopy (Murray et al. 1984). Microbial Mn oxides include hausmannite (Mn3O4), fietknechite (β-MnOOH), manganite (γ-MnOOH) (Mandernack et al. 1995b), todokorite (Takematsu et al. 1988), and γ-MnO2 (Greene and Madgwick 1991) (Table 3). In pure bacterial cultures, Mn is oxidized initially to a low valence state, predominantly hausmannite, that later disproportionates to MnO2 (Hem and Lind 1983; deVrind et al. 1986) in a two-step process: > Mn3O4 + 6H+

3 Mn2+ + 3H20 + 1/2 O2 Mn3O4 + 4 H+

> MnO2 + 2 Mn2+ + 2 H2O

Scanning electron microscopy (SEM) studies have showed that microbial Mn oxides are not highly crystalline and are amorphous, recrumpled, and sheety microcrystalline solids. Prolonged incubation of cultures for a few weeks or months resulted in more crystalline forms of Mn oxides (Mandernack et al. 1995b). However, Murray et al. (1985) reported that even after 8 mon the reaction did not proceed beyond γ-MnOOH. The reaction was also rapid in initial stages and, once the cells and sheaths were covered with an excess of Mn oxides, autoxidation predominated over bacterial oxidation. The morphology of Mn oxides formed in pure cultures of bacteria was very similar to that of many naturally occurring Mn precipitates surrounding microbes in environmental samples (Ghiorse 1984b). Factors Governing Mn Oxidation and Oxides Formed The type of oxide formed can vary according to the type of microorganisms and changes in chemical, physical, and growth conditions of cultures (see Table 3). Metallogenium cultures catalyzed the deposition of disordered Mn oxides such as vernadite (δ-MnO2) in association with a Mn-oxidizing fungus (Emerson et al. 1982). In general, low Mn oxide concentration (nM–µM) and low temperature promoted direct oxidation of Mn(II) to Mn(IV) (Table 4) without any intermediate steps, as reported earlier by Rosson and Nealson (1982).

Table 3. Manganese oxides formed by microorganisms. Organisms

Mn oxides formed

Reference

Bacillus sp. SG1 Leptothrix sp. Pseudomonas sp. Pseudomonas in association with Chlamydomonas sp. Pedomicrobium sp.

Hausmannite, manganite Hausmannite, manganite Manganite Manganite γ-MnO2

Mandernack et al. 1995b Adams and Ghiorse 1988 Nealson et al. 1988 Greene and Madgwick 1991

Manganite

Larsen et al. 1999

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Table 4. Cr(VI)-tolerant levels of selected bacteria.

Organism Pseudomonas K21 Pseudomonas fluorescens Arthrobacter sp. Agrobacterium sp. Escherichia coli Frankia strains

Tolerable Cr(VI) concentration (mg L−1) 5356 >400 450 100 66 52–91

References Shimada 1979 Bopp 1980 Coleman and Paran 1983 Coleman and Paran 1983 Thompson and Watling 1984 Richards et al. 2002

Mn Oxides in Soil In soils, the commonly occurring Mn oxides are birnessite and vernadite (McKenzie 1989). Birnessite is mainly formed in neutral and alkaline soils whereas in acid soils the coprecipitates may be manganites. In flooded soils, the intermediate oxides Mn(Fe)2O3 (bixbyite), 3(MnFe)2O3 ⴢ MnSiO3 (braunite), (Mn+2Fe)(Mn+3Fe)2O4 (jacobsite), and Mn3O4 ⴢ Fe3O4 (vrendenburgite), and perhaps γ-MnOOH (manganite) and Mn3O4 (hausmannite), may be present. When the flooded soils are drained, coprecipitates of iron and manganese are probably formed (Ponnamperuma et al. 1969). Most of the Mn oxides present in soil, especially manganate and birnessite (the highly reactive forms of Mn oxides in Cr oxidation), are also produced by microbial cultures (Section II.E). Biogenic Mn Oxides Responsible for Cr(VI) in Long-Term Tannery WasteContaminated Soil In long-term tannery waste-contaminated soil at the Mount Barker site in South Australia, Cr(VI) was detected in surface and subsurface soil samples and in groundwater water samples (collected in piezometers below 50 cm) at levels far above the permissible level of 0.05 mg (kg−1 or L−1), even 25 yrs after last disposal of the tannery waste to the site. Water-soluble Cr(VI) in the surface soil samples collected in 1997–1998 was about 4 mg kg−1 soil (Naidu et al. 2000b). Even after 4 or 5 yrs, surface soil samples contained >3.5 mg Cr(VI) kg−1 soil (Kamaludeen 2002). It appeared that there was no appreciable attenuation of Cr(VI) at this site with time. This finding was surprising because the highly contaminated soil at the waste disposal site was rich in organic carbon (15.7%), derived from animal hides containing electron donors that normally should enhance the reduction and thereby also the attenuation of Cr(VI). The contaminated soil contained 0.3–0.6 mg Mn kg−1, mainly as insoluble Mn oxides, and a high population of Mn oxidizers (4.7–5.4 × 103 colonyforming units, CFU) in the surface soil (Kamaludeen 2002). A close correlation existed between concentration of total Mn in the soil and the concentration of Cr(VI) in soil solution. Although Mn oxidation in soil is known to be essentially mediated by microrganisms, there was no convincing evidence to suggest that biologically produced Mn oxides catalyzed the formation of Cr(VI) in the contaminated soil. In a most recent study, Kamaludeen

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(2002) found that Mn-enrichment cultures, prepared from long-term tannery waste-contaminated soil at the Mount Barker site, and a bacterium isolated from this enrichment culture, could oxidize the Mn(II) to Mn(IV) that, in turn, oxidized Cr(III) to Cr(VI). This finding provided probably the first evidence for the involvement of Mn-oxidizing bacteria in the oxidation of Cr(III) to Cr(VI) at the long-term contaminated site. Such biotic–abiotic coupling leading to the oxidation of Cr(VI) would explain why natural attenuation of Cr(VI) is not taking place at the Mount Barker site even in the presence of electron donors that should enhance natural attenuation.

IV. Implications of Chromium Transformations on Microorganisms and their Activities Generally, Cr exists in the environment in stable oxidation states, Cr(III) and Cr(VI). The effects of Cr on microorganisms are governed by its speciation. Cr(VI), a strong oxidant with a high solubility in water, is distinctly more toxic than relatively less soluble Cr(III). In alkaline soils, Cr(VI) in solution is the dominant species. In soils at around pH 5.5, however, Cr(VI) complexes with organic matter and is reduced to Cr(III). Cr(III) is then readily precipitated in acidic soil as insoluble oxides and hydroxides and is hence less bioavailable and less ecotoxic than Cr(VI). The impact of heavy metals such as cadmium, lead, copper, and zinc on microorganisms and their activities in soils has been more extensively studied (Doelmann and Haanstra 1986; Vig et al. 2002) than that of Cr. Previous research on the toxicological effects of heavy metals has focused mostly with soils contaminated over a long term with sewage sludge generally containing multimetals and complex organic contaminants and to some extent with soils freshly spiked with individual or mixed metals. Short-term exposure to contaminants as in freshly spiked soils causes acute, probably temporary, toxicity. In long-term contaminated soils, chronic effects of the pollutants with a persistent shift in microbial populations can be common as a result of elimination of sensitive microorganisms coupled with selective stimulation of microorganisms that are already tolerant or resistant to the pollutant or have evolved by adaptation over a long duration of exposure. A. Microorganisms Chromium(VI), which can easily diffuse across the cell membrane in prokaryotic and eukaryotic organisms, is reduced to Cr(III) inside the cells with some reports on transitory formation of Cr(V) and Cr(IV) as toxic intermediates (Arslan et al. 1987; Liu et al. 1995). On the other hand, entry of Cr(III) into the cells is restricted due to its precipitation as hydroxides. Hence, Cr(VI) is more toxic to microorganisms than Cr(III) (DeFlora et al. 1984). Many microbial cells showed negative response on contact with Cr. Genotoxic effects in microbial cells are mainly impacted by Cr(VI), resulting in frameshift mutations (Petrilli

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and deFlora 1977) and lethal DNA damage (Ogawa et al. 1989). Accumulation of Cr resulting in cell sequestering to combat its inhibitory effect is a major phenomenon in many microbes resistant to Cr. Prokaryotes are more resistant to Cr than are eukaryotes. Thus, microbes exhibit different Cr tolerance levels (see Table 4). Some of the important changes in microorganisms caused by Cr are summarized in Table 5. Fungi Cr–fungi interactions, mostly related to chromate resistance in filamentous fungi and yeasts and chromate reduction by yeasts, have been extensively studied (Cervantes et al. 2001). Generally, fungi are less sensitive than bacteria to metals (Doelman 1985). Cr compounds at environmentally relevant concentrations (Naguib et al. 1984), however, inhibited the tomato pathogenic fungi, Fusarium oxysporum f. sp. lycopersici and Cunninghamella echinulata. Fungi, mostly yeasts, with a high degree of resistance to chromate have been isolated from Cr-contaminated [long-term or short-term (spiked with Cr)] soil environments. Laboratory induction through enrichment in Cr-amended soil suspensions or repeated transfers of fungal strains in Cr-supplemented media and mutagenesis by UV irradiation or chemical mutagens have also been used for isolation of chromate-resistant fungi. Yeasts resistant to chromate include Candida albicans, Schizosaccharomyces pombe, and Saccharomyces cerevisiae. Chromate resistance in fungi may have several reasons: decreased uptake of Cr(VI) (CzakoVer et al. 1999), defect in sulfate transport (Lachance and Pang 1997), involvement of vacuolar structures (Gharieb and Gadd 1998), reduction of Cr(VI), and presence of acid phosphatase (Raman et al. 2002). Decreased uptake of Cr(VI) is the major mechanism of chromate resistance in both filamentous fungi and yeasts. Biological reduction of Cr(VI), as in many bacteria, can also be important in the chromate resistance of yeasts, but not in filamentous fungi. Mycorrhizal Fungi There is considerable interest in exploiting the potential of mycorrhizal fungi in afforestation and reclamation of degraded lands, mine spoils, and metal-polluted soils. For instance, trees inoculated with ectomycorrhizal fungi become established better than trees without mycorrhizal association in metal-polluted soils (Brown and Wilkins 1985; Jones and Hutchinson 1986). Metal-tolerant mycorrhizal fungal strains, developed in the laboratory by repeated transfers in a metal-containing medium and then used for inoculation, have an advantage over sensitive strains in forming an effective association with host trees in metal-polluted soils. A survey on the distribution of vesiculararbuscular mycorrhizal (VAM) fungi in tannery waste-polluted soils at three sites in Tamil Nadu, India (Raman and Sambandan 1998) revealed the occurrence of 15 species of VAM fungi in the polluted soils. Of the 22 plant species from the polluted sites screened, 19 plant species harbored a mycorrhizal association. Glomus fasciculatum, G. geosporum, and Gigaspora gigantea were the dominant VAM fungi in the tannery waste-polluted soils with a high concentration of total Cr [1400–1800 mg/kg soil; Cr(VI) level not provided]. The trees

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Table 5. Effect of chromium on microorganisms. Organism

Cr species and concentration

Effects

Reference

Staphylococcus aureus, S. epidermis, Bacillus cereus, B. subtilis

1040 mg Cr(VI)/L

Bondarenko and Ctarodoobova 1981

Shigella sonnei, Shigella flexneri, Salmonella typhosa, Proteus mirabilis, and Escherichia coli E. coli Arthrobacter sp.

3027 mg Cr(VI)/L

Small colonies, cell elongation, cell enlargement, results in filamentous forms Small colonies

Photobacterium phosphoreum Algae Euglena gracilis

Cr(VI) twice toxic than Cr(III)

Euglena gracilis Scenedesmus sp. and Selenastrum

Cr(III) Algal growth inhibited by Cr(VI), and not by Cr(III), at 100 mg/L Growth not inhibited by Cr(III) or Cr(VI) at 45–100 mg/L No growth above 15 mg Cr/L

Chlorella vulgaris Scenedesmus acutus MGT: mean generation time.

Cr(VI)

Filamentous forms observed Depressed growth rate; MGT increased from 2.1 to 10.2 hr MGT increased from 1 to 3 hr Inhibits prodigiosin production Utilizes chromates and dichromates; results in clumping of cells Reduction in light emission Lag growth phase increased, arrest of cells in G-2 phase; inhibition of photosynthesis and respiration Growth inhibited

Theotou et al. 1976 Coleman and Paran 1983 Coleman and Paran 1983 Furman et al. 1984 Romanenko and Korenkov 1977

Qureshi et al. 1984 Brochiero et al. 1984

Brochiero et al. 1984 Brady et al. 1994

Travieso et al. 1999 Travieso et al. 1999

S.P.B. Kamaludeen et al.

Agrobacterium sp. Serratia marcescens Pseudomonas dechromaticans

Cr(VI) Cr(VI) 0–200 mg/L 0–100 mg/L 19 mg Cr(VI)/L

Bondarenko and Ctarodoobova 1981

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Prosopis juliflora and Azadirachta indica, which grew well at the polluted sites, harbored a high density of VAM fungi. Acid phosphatase activity, extracellular activity in particular, has been implicated in imparting heavy metal resistance to mycorrhizal fungi. The resistance mechanism involved the precipitation of the metals by HPO4 released by acid phosphatase, followed by the binding of the precipitated metal to the cell surface. Of the ectomycorrhizal fungi, Laccaria laccata and Suillus bovinus, the former, which produced more acid phosphatase, was more tolerant to high concentrations of Cr(VI) (Raman et al. 2002). The compatibility of mycorrhizal fungi with the host plant together with their tolerance to the metal would dictate the successful establishment of the plant in metal-polluted environments. Algae The interactions between Cr and algae, terrestrial or aquatic, have been less intensively studied than Cr–bacteria and Cr–fungi interactions as is the case with other metal and organic pollutants. There are reports of tolerance or resistance of a limited number of algae to Cr, depending on its speciation, but the mechanism(s) involved in algal resistance are not understood. Reduction of Cr(VI) to Cr(III) and decreased uptake of Cr by algal cells are not probably involved in algal resistance to Cr. Sequestration of Cr by its complexation with organic compounds in algal exudates is a possibility, but needs confirmation. Interference in sexual reproduction has been implicated in the evolution of Crtolerant algae (Corradi et al. 1995). Recently, Megharaj (unpublished data) found total suppression of algal growth in a long-term tannery waste contaminated soil with high levels of Cr [total Cr, 62,000 mg/kg; Cr(VI), 40 mg/kg], when contaminated soil was incubated under moist conditions for 6 mon or more. Viti and Giovannetti (2001) examined the impact of Cr concentration on photosynthetic microorganisms in three soils whose Cr concentration ranged from 67 to 5490 mg/kg. Chronically high concentrations of Cr adversely affected aerobic photosynthetic microorganisms and drastically reduced the population (by most probable number technique) of nitrogen-fixing cyanobacteria. Soils polluted with Cr harbored a low population of the cyanobacteria of the genus Nostoc, and rarely with heterocysts. In soil enrichment cultures with low Cr levels, however, Nostoc dominated and possessed numerous heterocysts. Toxicity of Cr to algae may involve not only Cr(VI) but also Cr(V), because transitory accumulation of toxic Cr(V) has been reported in algal cultures of Spirogyra and Mougeotia (Liu et al. 1995). In a study on the impact of Cr on algae, total algal counts should be complemented by changes in algal biodiversity. Cr-tolerant algal populations increased in river waters receiving toxic levels of Cr from a paper mill (Sudhakar et al. 1991). It is not always possible to determine the impact of a metal on the changes in algal biodiversity in soil or water environments because of the practical difficulty in selecting an appropriate control under field situations, for instance, adjacent to long-term contaminated polluted sites. Garnham and Green (1995) studied the accumulation of chromate ions by a unicellular non-nitrogen-fixing cyanobacterium, Synechococcus sp. PCC 6301,

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and a filamentous nitrogen-fixing cyanobacterium with heterocysts, Anabaena variabilis, and their ability to reduce Cr(VI). Both cyanobacteria accumulated chromate in the cell walls rapidly, but at a low level, depending on its concentration; biosorption was an energy-independent process. Cyanobacteria are known to produce and release complex organic ligands that can bind metals (Megharaj et al. 2002). During 18-d growth, A. variabilis reduced almost all the added chromate to Cr(III) in stoichiometric amounts, with 50% of the latter in the cells and remaining 50% in the medium (Garnham and Green 1995). Synechococcus sp. PCC 6301 was unable to reduce Cr(VI). Cr(VI) reduction by A. variabilis presumably occurred in the heterocysts. It may be worthwhile to mention that chromate reduction by A. variabilis proceeded at a slow rate when compared to that reported in bacterial cultures. Bacteria Gram-positive bacteria are more resistant to Cr than gram-negative bacteria (Ross et al. 1981). Chromium(VI), at 10–12 mg/mL, inhibited most soil bacteria in liquid media whereas Cr(III) at this concentration was not toxic. Pilz (1986) found that the toxicity of Cr(VI) in aqueous media differed with bacterial strains used, and EC50 values for Cr(VI) ranged from 0.003 to 7000 mg/L. Likewise, Cr was toxic to mixed bacterial populations of sewage origin in a chemostat (Lester et al. 1979). Hexavalent Cr is toxic and mutagenic to most organisms including algae and bacteria (Wong and Trevors 1988). In a more recent study (Francisco et al. 2002) using sodium dodecyl sulfate-polyacrylamide gel electrophoresis (SDS-PAGE) protein patterns and fatty acid methyl ester analysis, the major group of bacteria, isolated from a Cr-contaminated activated sludge with total Cr level of 1.3%, belonged to γ-Proteobacteria, exclusively with strains from the genus Acinetobacter. Evidence suggested that the presence of Cr(VI) had no effect on the viability of γ-Proteobacteria. Hattori (1992) applied Cr at 10 µmol/g as CrCl3 and other heavy metals to two soils, and 3 d later the soils were amended with 2% sewage sludge. The inhibitory effect of Cr(III) on bacterial population was more pronounced in Gley soil than in light-colored Andosol soil. High toxicity of Cr(III) in Gley soil was associated with increased bioavailability (water soluble and CaCl2 extractable). Concomitant with inhibition of bacteria in Cr-treated soil samples, the fungal population increased severalfold over that in control soils not treated with Cr. With regard to metal toxicity, in terms of ED50 (concentration at which the number of bacterial colonies from soil dilutions plated on a medium decreased to 50% of that in control), Cr (ED50 > 100 µmol) was distinctly less toxic than Cd (ED50 < 20 µmol). B. Effects on Soil Microbial Community Phospholipid fatty acids (PLFA) are a good indicator of environmental disturbance. The principle is based upon the fact that different subsets of a microbial community differ in their fatty acid composition. Membrane lipids and their

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associated fatty acids are particularly useful biomarkers as they are essential components of every living cell and have great structural diversity, coupled with high biological specificity. Also, by using the phospholipid composition only, one can ensure that the measurement is on the living part of the microflora, because phospholipids are assumed to decompose quickly after the organism dies. The PLFA pattern can therefore be viewed as an integral measurement of all living organisms present in that sample, reflecting both species composition and relative species abundance (Baath 1989). Phospholipid fatty acids have been useful in distinguishing the abundance and structure of microbial communities in soils (Zelles 1999). There are several reports on the shift in microbial populations in soils contaminated over the short and long term with Cu, Pb, Zn, and Ni as compared to uncontaminated soils (Frostega˚rd et al. 1993; Pennanen et al. 1996; Griffiths et al. 1997; Kelly et al. 1999a,b). In most cases, the multivariate principal component analysis (PCA) differentiates the PLFA patterns of contaminated soils from those of uncontaminated soils. PCA of PLFA profiles indicated distinct decreases in fatty acids specific for certain microbial populations such as actinomycetes (18 : 0 10 Me), VAM fungi (16 : 1 ω5c), and other fungi (18:2 ω6c; 20:2 ω6c) even after 18 yr of amendment with dewatered sewage sludge containing multimetals, including Cr [44.5 Cd, 512 total Cr, 341 Cu, 159 Ni, 337 Pb, and 1506 Zn in mg kg−1; Cr(VI) not estimated] (Kelly et al. 1999a). Such relative decreases in several fatty acids in sludge-amended soils suggested inhibition of several specific populations of soil microorganisms. However, in sludge-amended soils, counts of culturable bacteria significantly increased, in contrast to >20-fold decrease in dehydrogenase activity (DHA). There has been no report hitherto, however, on PLFA patterns and shift in microbial populations in soils freshly spiked with Cr alone or in long-term tannery waste-contaminated soils with high levels of Cr as the major pollutant. Past studies have shown that chronic metal stress affects the microbial community and decreases microbial biomass, activity, and diversity. Most studies on microbial community structure under chronic metal stress, however, have been confined to soils treated with sewage sludge-containing multimetals, often with Cr as a minor constituent. Francisco et al. (2002) attempted to establish a relationship between a culturable microbial community [characterized by fatty acid methyl ester (FAME) analysis and SDS-PAGE] and the Cr(VI) resistance and Cr(VI) reduction ability of the representative strains of each population in activated sludge (total Cr, 0.197–2.5 ng/L), under chronic Cr stress, from urban and industrial tannery areas in Portugal. The ability for aerobic reduction of Cr(VI), when examined with 28 strains representative of each FAME cluster and noncluster in nutrient broth containing 1 mmol/L, was not restricted to one species or one genus and was widespread in several Proteobacteria subclasses and in gram-positive G + C bacteria. Cr(VI) resistance and Cr(VI) reduction are not exclusive to a single group, possibly a result of horizontal genetic transfer under selective pressure from chronic Cr contamination. In bioremediation strat-

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egies using microbially mediated Cr(VI) reduction, there is a need for a better understanding of the microbial community and the population response under chronic metal stress. C. Effects on Soil Microbial Processes and Activities Usually, concentration of Cr in soil varies from 100 to 300 mg kg−1; however, the concentration of Cr available to soil microflora is low. The toxic effects of Cr are mainly governed by speciation and bioavailability rather than by the total Cr concentration. In alkaline soils, Cr(VI) in solution is dominant, resulting in increased inhibition, but in acidic soils most of the Cr(VI) complexes with organic matter and is reduced to Cr(III), leading to a decrease in toxicity. Microbial Biomass The effect of heavy metals, with Cr as one of the major contaminants, on microbial biomass has been studied at grassland sites receiving multimetal-rich military waste disposals (Kuperman and Carreiro 1997) or treated with timber preservatives containing Cu, Cr, and As (Bardgett et al. 1994). Combined concentrations of heavy metals distinctly suppressed the microbial biomass (fungal and bacterial) in grassland sites receiving military wastes (Cr, 42–143 mg/kg) or wood preservatives. In soils polluted with military wastes, total and fluorescein diacetate-(FDA-)active fungal biomass was more sensitive than FDA-active bacterial biomass (Kuperman and Carreiro 1997). It was not clear whether the inhibitory effect on microbial biomass in soils polluted with military wastes was caused by heavy metals or increased soil pH. Likewise, Bardgett et al. (1994) reported a greater sensitivity of fungal biomass, relative to bacterial biomass, in pasture soils polluted with timber preservatives. In these studies with multimetals, however, it is difficult to distinguish the individual effect of Cr. Dehydrogenase Activity (DHA) Dehydrogenases are essential enzymes, involved in oxidoreduction processes, in all microorganisms. Dehydrogenase activity (DHA) is one of the important parameters widely used to study the ecotoxic effects of metals and organic contaminants. The main advantage of this simple, but sensitive toxicity assay is that it reflects the overall microbial activity of the active microbial populations in the soil to provide the current status of soil health. Reports on the effect of Cr on microbial processes of importance to soil fertility are summarized in Table 6. Generally, DHA decreases in sewage sludge-amended soils. Because sewage sludge contains a mixture of heavy metals and organic contaminants, it is difficult to identify the metal or the contaminant responsible for the specific effects. In contrast, an increase (18%–25%) in DHA has also been reported in soils amended with sewage sludge containing 220 µg Cr g−1 of soil. This increase was more pronounced in sandy loam than in loam or clay loam soils (Dar 1996). Soil factors such as pH, moisture content, and cation-exchange capacity (CEC) (Doelmann and Haanstra 1979) influence the DHA in soils. Soil pH determines

Table 6. Effect of chromium on microbial processes in soils and pure cultures of microorganisms. Soil type

8.6

Sandy loam pH 6.6, 0.84% C

Chang and Broadbent 1981

400

—, CO2, 17%

Doelmann 1985

260 26 50

—, CO2, 15% —, CO2, 10% —, CO2, 20%

Lighthart et al. 1983 Lighthart et al. 1983 Skujins et al. 1986

0, N mineralization

Ruhling and Tyler 1973

400

—, N mineralization, 40%

Chang and Broadbent 1982

100 260

—, nitrification, 40% —, N mineralization, 18% —, nitrification

Liang and Tabatabai 1977

269

—, nitrification, 81%

Liang and Tabatabai 1978

200

—, nitrification, 26%

Skujins et al. 1986

0.32Cd + 5.6Cu + 7.2Pb + 30.7Zn + 0.76Cr 2.24Cd + 7.5Cu + 47Pb + 148Zn + 15.7Cr 0.56Cd + 1.88Cu + 12Pb + 37Zn + 3.9Cr 200 Cr + Ni + Mo

—, N fixation, 93% 0, nitrification

Skujins et al. 1986 Wilson 1977

—, nitrification

Wilson 1977

+, lag nitrification

Wilson 1977

—, phosphatase, 20%

Ruhling and Tyler 1973

50

133

Forest humus pH 3.6–4.1

—, CO2, 10%

10000

Soil litter, agricultural pH 5.8–7.8; 2.6%–5.5% C Agricultural, pH 5.8–7.8 2.6%–5.5% C Forest sandy clay pH 7.0

References

Chromium–Microorganism Interactions

Silt loam 1.31% C+ 1% dry sludge and 1% alfalfa Sandy pH 7.0, 1.6% OM, and CEC 1–2 pH 6.2, 64% OM pH 6.7, 3.1% OM Forest sandy loam pH 7.0 Forest humus pH 3–4.2 Silt loam pH 6.9 + 1% sludge and alfalfa

Effects, parameter measured, % inhibition

Cr concentration (µg/g)

134

Table 6. (Continued).

Soil type

1500

—, dehydrogenase, 40%

7500

0, dehydrogenase

7500

0, dehydrogenase

Doelmann and Haanstra 1979

390–1880

—, urease, 10%

360

— urease, 10%

150

—, dehydrogenase, 83%

Rogers and Li 1985

200

—, urease, 28%

Skujins et al. 1986

269

—, urease, 17%–50%

Tabatabai 1977

—, CFU bacteria 20%

Zibilske and Wagner 1982

1.0 1.0 556 55, 150, 400, 1000, 3000, 8000, as CrCl3 130 as CrCl3 55–2000 as CrCl3

References

—, ATP, 60% 0, altered fungal community Toxic to arylsulfatase at 6 weeks; decreased at 18 mon —, arylsulfatase, 41% (average for four soils —, soil respiration (glutamic acid as substrate) 23% in sand; 5.12% in clay

Doelmann and Haanstra 1986

Haanstra and Doelman 1991 Al-Khafaji and Tabatabai 1979 Haanstra and Doelman 1984

S.P.B. Kamaludeen et al.

Sandy pH 5.7, 2.8% OM Clay pH 7.5, 3.2% OM Peat pH 5.7, 46% OM Sandy pH 7.0, 1.6% OM Sandy peat pH 4.4, 12.8% OM Agricultural soil 1.3% OM Forest sandy loam pH 7.0 Six soils (pH 5.1–7.8; 1.5%–5.5% C) Silt loam pH 6.0, 2.1% OM

Five soils (pH 4.4–7.7; 1.6%– 12.8% OM) Four soils (pH 6.2–7.0; 2.7%– 5.3% C) Five soils (pH 4.4–7.7; 1.6%– 12.8% OM)

Effects, parameter measured, % inhibition

Cr concentration (µg/g)

Table 6. (Continued). Soil type

Cr concentration (µg/g)

Effects, parameter measured, % inhibition

References

Two soils (pH 5.9 and 6.4)

Cr(III): 100 Cr(VI): 10 and 100

Four soils (pH 5.6–7.6; 2.6%– 4.7% C) Three soils (pH 5.6–7.6; 2.6%– 4.7% C) Three soils (pH 5.8–7.4; 2.6%– 5.5% C)

260 (CrCl3)

—, soil respiration at 3000 µg/ g, + at 8000 µg/g —, 35% in Gley soil, 5% in Andosol; degree of inhibition related to water-soluble Cr —, CO2 —, CO2 —, bacteria, Cr(VI) > Cr(III) —, L-asparaginase, 12%–21%

260 (CrCl3)

0, amidase

Frankenberger and Tabatabai 1981

130 as CrCl3

—, acid phosphatase 27% —, alkaline phosphatase 33% —, acid phosphatase 5% —, alkaline phosphatase 14% —, DHA (significant inhibition with increasing leached soil concentrations of preservative constituents (fluoride and total Cr); rye meal supplement led to increased levels of DHA in contaminated soils Toxicity to DHA: based on total dose (ED50): Hg > CU > Cr(VI) > Cr(III) > Cd > Pb; based on solution concentration (EC50): Hg > Pb > Cu > Cd > Cr(III) > Cr(VI)

Juma and Tabatabai 1977

Five soils (pH 4.4–7.7; 1.6%– 12.8% OM) Two soils (pH 5.8 and 6.4; 0.5% and 3.2% C)

55–8000 as CrCl3 520 CrCl3

Sandy loam soil (pH 6.15; organic matter 33.6%)

Adjacent to remedially treated (with chromate fluoride wood preservative) timber poles

Loess soil (pH 7.02; 1.12% C; 15.2% clay

Cr(III) as nitrate and in the oxyanion Cr(VI) as a K salt; 8–12 geometrically increasing doses

Hattori 1992 Ross et al. 1981 Frankenberger and Tabatabai 1991

Sinclair et al. 1997

Chromium–Microorganism Interactions

13 as CrCl3

Doelman and Haanstra 1984

Welp 1999

135

136

Table 6. (Continued).

Soil type Chromium-contaminated activated sludge (chronic stress)

References

Total Cr, 0.197–2.5 ng/L

Attempted to establish a relationship between culturable microbial community (FAME analysis and SDS-PAGE) and Cr(VI) resistance and reduction under chronic Cr stress

Francisco et al. 2002

CrO3 (7.8, 15.6, 23.4, 39.0, and 47.0

In vitro growth of both fungi stimulated at 0.15 mM Cr but inhibited at 0.15 mM; acid phosphatase activity of both fungi increased at all Cr concentrations; alkaline phosphatase increased in L. laccata and decreased in S. bovinus because of Cr Growth inhibited by Cr(VI), but only at 52 mg/L Growth and Mn peroxidase activity inhibited by Cr(VI) at 10 mg/L; no toxicity to Mn peroxidase on exposure to Cr(VI) after lag phase of mycelial growth

Raman et al. 2002

Stereum hirsutum Polyporus ciliatus, Stereum hirsutum

Effects, parameter measured, % inhibition

Cr(VI) as K2Cr2O7 from 5 to 20 mg/L

OM: organic matter; CFU: cology-forming units; DHA: dehydrogenase activity.

Baldrian and Gabriel 1997 Yonni et al. 2002

S.P.B. Kamaludeen et al.

Fungi Laccaria laccata, Suillus bovinus (ectomycorrhizal fungi)

Cr concentration (µg/g)

Chromium–Microorganism Interactions

137

the amount of metal available to the microbes in soil solution and thereby its eventual effect on DHA. Likewise, moisture at field capacity might mask the effect of heavy metals on DHA. Long-term incubation of soils with heavy metals may have a great impact on DHA. However, in general, Brendecke et al. (1993) found that DHA and soil respiration were little affected by sewage sludge containing multimetals even after 4 yrs of its application. However, Kelly et al. (1999a) reported the inhibition of DHA in a soil 18 yr after application of a sludge containing multimetals including total Cr (512 mg/kg). A decreased toxicity of the metals was observed in most cases as the exposure time increased, which was attributed to the elimination of the sensitive microbial populations by the chronic effects of the heavy metals with a concomitant shift toward the dominance of tolerant microorganisms. The increased abundance of tolerant organisms in polluted environments can be caused by genetic changes, physiological adaptations, or replacement of metal-sensitive species with species that already are tolerant of that heavy metal. Bacterial cultures, e.g., Pseudomonas, could tolerate maximum Cr(VI) concentrations of about 5356 mg L−1. Thus, a distinct shift in population can occur in contaminated soils, especially under long-term impact. Several techniques such as phospholipid fatty acid (PLFAs) and denaturing gradient gel electrophoresis (DGGE) have been used recently to determine the microbial populations in agricultural soils and in soils polluted with organics and inorganics. Among these techniques, PLFA has been used widely to determine the impact of metals on microbial communities in soils. Wood preservatives can be a serious source of environmental contaminants. For instance, softwood timbers are often treated with a preservative containing Cr, Cu, and As as protection against insect and fungal attack. The effects of surface runoff from such a treatment plant on biological activities in pasture soils with low, medium, and high levels of contamination were examined by Yeates et al. (1994). Metal content in the soil samples ranged between 47 and 739 mg Cr kg−1, 19 and 835 mg Cu kg−1, and 12 and 790 mg As kg−1. Highly contaminated soil samples in the surface layer contained at least 700 mg each of Cr, Cu, and As kg−1. Generally, normal microbial processes (DHA, basal respiration, substrate-induced respiration, nitrification, phosphatase activity) were initially inhibited, especially at higher levels of contamination by the three heavy metals, DHA was the only activity that was distinctly inhibited even after 6 wk. In another instance where creosoted electric poles were treated with chromated fluoride wood preservative to eradicate basidiomycete fungi, negative effects on soil DHA were associated with increased soil concentrations of leached fluoride (160–960 mg/kg) and total Cr concentrations (74–218 mg/kg) (Sinclair et al. 1997). Application of rye meal largely alleviated the toxicity of preservative pollutants on DHA. There are not many studies on the relative toxicity of Cr species, Cr(III), and Cr(VI) and their toxicity in relation to other metals to the microbial activities in soil. Welp (1999) found that, based on total dose, the toxicity [ED50 values (mg/ kg) given in parentheses] of heavy metals to soil DHA in a loess soil decreased

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S.P.B. Kamaludeen et al.

in the order Hg (2.0) > Cu (35) > Cr (VI) (71) > Cr(III) (75) > Cd (90). Sorption and solubility data, however, revealed that Cr(VI) was the least sorbed and yet least toxic to DHA among the metals tested, including Cr(III). Based on solution concentrations of metals, the toxicity to DHA, in terms of EC50, followed the order Hg (0.003) > Cu(0.05) > Cd (0.14) > Cr(III) (0.62) > Cr(VI) (78.1). One would expect that least sorbed Cr(VI) with high solubility is more bioavailable and hence more toxic to microbial activities than other metals. Surprisingly, based on solution concentration, Cr(III) appeared to be more toxic than Cr(VI), contrary to the common notion, which is difficult to explain. In a long-term field experiment, application of high and low rates (0, 30, 90, and 270 t/ha) of municipal waste composts, containing multimetals including Cr (total Cr, 31 mg/kg; available Cr, 0.7 mg/kg), had no inhibitory effect on various soil enzyme activities (alkaline phosphomonoesterase, phosphodiesterase, arylsulfatase, dehydrogenase, and L-asparaginase) 3 yr after their application. In fact, these enzyme activities increased linearly up to 90 t/ha (Giusquiani et al. 1994). Other Enzymes Among the 21 trace elements (applied at 1300 mg/kg) screened for toxicity to arylsulfatase activity, average inhibition of the enzyme activity by Cr, applied as CrCl3, in four soils used was 41% over control (AlKhafaji and Tabatabai 1979). Cr(III) appeared to be the most toxic to enzyme activity, when assayed within 30 min after metal addition, among the heavy metals screened as follows: Cr > Cd > Zn > Cu ≥ Ni > Pb. Its inhibitory effect decreased by 10 fold when the metal concentration decreased from 1300 to 114 mg/kg. Cr(III) was less inhibitory than Ag(I) and Hg(II). Haanstra and Doelman (1991) found that toxicity of Cr (applied as chloride at 0–8000 mg/kg) to arylsulfatase activity decreased with time between 6 wk and 18 mon. After 18 mon, Cr was the least toxic among the metals screened. Likewise, Cr(III) applied at 260 mg/kg inhibited L-asparaginase (Frankenberger and Tabatabai 1991) and phosphatase (Juma and Tabatabai 1977) activities. Soil amidase (Frankenberger and Tabatabai 1981) was less affected by Cr and other heavy metals than urease, arylsulfatase, and phosphatase. The activities of N-acetylglucosaminidase, β-glucosidase, endocellulase, and acid and alkaline phosphatases, as were fungal and bacterial biomass, were less pronounced in grassland soils polluted with military wastes containing multimetals including Cr than in reference soil (Kuperman and Carreiro 1997). Doelman and Haanstra (1989) developed an ecological dose–response model using phosphatase activity to determine the short- and long-term effects of heavy metals. Cocontamination of polluted soils and wastewaters with recalcitrant aromatic compounds and heavy metals can be common and widespread. White rot fungi are known for their ability to mediate the degradation of several recalcitrant aromatic compounds. Lignolytic activity of these fungi is catalyzed by extracellular oxidative enzymes such as laccase and manganese peroxidase. Heavy metal-tolerant white rot fungi would have a better scope for remediation of aromatic compounds in a cocontaminated site. The effect of heavy metals

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139

[Cr(VI), Cd(II), Zn(II), Pb(II), and Ni(II)] on the growth and manganese peroxidase activity of two wood-rotting Basidiomycetes, Polyporus ciliatus and Stereum hirsutum, has been reported (Yonni et al. 2002). Cr(VI) was inhibitory to the growth of both these white rot fungi at concentrations of 10 µg/mL and above, and growth was totally inhibited at 20 µg/mL. According to an earlier report (Baldrian and Gabriel 1997), mycelial growth of S. hirsutum was likewise inhibited by Cr(VI), but only at a high concentration of 1 mM. The manganese peroxidase activity of both Polyporus ciliatus and Stereum hirsutum was inhibited by Cr(VI) (10 µg/mL), Pb(II) (5 and 10 µg/mL), and Ni(II) (5 and 10 µg/ mL) although Cd(II) and Zn(II) were not inhibitory even at 20 µg/mL (Yonni et al. 2002). In combinations of Cr(VI) with one or two more heavy metals, added at individually subtoxic concentrations, growth of S. hirsutum was totally inhibited (Yonni et al. 2002). However, the toxic effect of Cr(VI) and other heavy metals, either individually or in combination, on growth and manganese peroxidase activity of S. hirsutum was alleviated when the metals were added after the lag growth of the fungus. This mechanism must be considered if S. hirsutum (probably other white rot fungi as well) is used for bioremediation of aromatic contaminants in environments polluted with heavy metals as cocontaminants. Nitrification Liang and Tabatabai (1978) found that the toxicity of metals (added at 5 mM/kg) to nitrification of NH+4-N followed the order Hg > Cr(III) > Cd > Ni > Cu > Zn > Pb, with an average inhibition >50%. The inhibitory effect of Cr(III) on nitrification was noticed in all the three soils used and ranged from 59% to 96%. Ross et al. (1981) suggested that Cr(VI) may impact nitrification in soils in view of its more pronounced toxicity even at low levels (1–10 µg/ mL) to gram-negative soil bacteria than gram-positive bacteria in liquid cultures. In a study on the effect of heavy metals on nitrogen transformations (N immobilization, N mineralization, and nitrification) in silt loam amended with NH+4-N (100 µg/g), 1% sewage sludge and 1% ground alfalfa (Medicago sativa), Cr was the most inhibitory to the N transformations (Chang and Broadbent 1982). At 400 µg/g, all metals inhibited the three N transformations. Inhibition of N transformations by heavy metals during a 2- to12-wk incubation followed the order Cr(III) > Cd > Cu > Zn > Mn > Pb. No clear relationship existed between the toxicity of heavy metals to N transformations and the metals extracted by water, KNO3, and diethylenetriaminopentaacetic acid (DTPA). Nitrogen Fixation Heterotrophic N2 fixation was sensitive to Cr at 50 mg/kg in soil spiked with Cr (Skujins et al. 1986). Likewise, nitrogenase of cyanobacteria was inhibited by 50% in soils treated with sewage sludge containing Cr (80 mg/kg soil) and five more metals at concentrations well below the European guidelines (Brookes et al. 1986). Soil Respiration Both Cr (III) (100 mg/kg) and CR(VI) (10 and 100 mg/kg) distinctly decreased the evolution of CO2 from two field-moist soils during

140

S.P.B. Kamaludeen et al.

3-wk incubation although Cr(III) was not toxic to bacterial growth in liquid cultures (Ross et al. 1981). Interestingly, phosphate (0.01 M KH2PO4-K2HPO4) extractable Cr(VI) decreased to 25% of the original level during this period. Yet, CO2 was not restored to that in the control soil indicating a persistent adverse effect of Cr(VI) on the microbial activities in the soil. Likewise, Hattori (1992) reported increased toxicity of Cr(III) (as CrCl3), applied at 10 µmol/g, to CO2 evolution in Gley soil over that in light-colored Andosol. In Cr-treated Gley soil, respiration was inhibited by 37% over that in the control. The increased toxicity to CO2 evolution in Gley soil was directly related to the increased bioavailability of water-soluble Cr. Substrate-induced respiration in soils is an active index of active microbial biomass. Mineralization of glutamic acid as substrate was used as a parameter for screening the toxicity of six heavy metals including Cr added as chlorides in six soils at concentrations ranging from 55 to 2000 mg/g (Haanstra and Doelman 1984). All six metals exerted the strongest inhibitory effect on glutamic acid mineralization in a sandy soil. Reports on long-term impact of individually applied Cr on microorganisms and their activities in soils are scant. Short-term (2, 4, and 8 wk) and long-term (18 mon) effects of Cr(III) and other heavy metals, applied as chlorides at 0, 55, 150, 400, 1000, 3000, and 4000 µg/g, on soil respiration were examined in five Dutch soils (Doelman and Haanstra 1984). In sandy loam, clay, and sandy peat soils, Cr distinctly inhibited soil respiration under both short-term and longterm incubation, irrespective of its concentration, although inhibitory effects partially decreased with time. In contrast, Cr stimulated soil respiration at unrealistically high concentrations of 8000 µg/g in sand and at 3000 and 8000 µg/g in silty loam soil. It is probable that Cr, as a trivalent cation, facilitated the increased availability of organic matter to microorganisms in these soils. Reductive Dechlorination of Organics Cocontamination of the soil and water environments with inorganic and organic contaminants by both natural sources and anthropogenic activities is a common occurrence worldwide. Kuo and Genthner (1996) reported the effect of sublethal and lethal concentrations of Cr(VI), Cd(II), Cu(II), or Hg(II) added at 0.01–100 µg/mL on the biotransformation of 2-chlorophenol, 3-chlorobenzoate, phenol, and benzoate in an anaerobic consortium. In general, these heavy metals extended the acclimation periods and distinctly retarded or totally inhibited the anaerobic dechlorination or biodegradation of the selected organic compounds, depending on their concentration. Among the metals used, Cr(VI) and Cd(II) were the most inhibitory to dechlorination of 3-chlorobenzoate in anaerobic consortium, with total inhibition at 0.5 µg/mL. Likewise, the dechlorination of 2-chlorophenol to phenol was inhibited by Cr(VI), but to a lesser extent than that of 3-chlorobenzoate, with total inhibition only at 5.0 µg/mL. Phenol, formed from 2-chlorophenol at sublethal concentration of Cr(VI), accumulated at sublethal concentration (2.5 µg/mL) of Cr(VI), but was biodegraded after acclimation. Biodegradation of added phenol and benzoate was inhibited. Although water-soluble Cr(VI) is 10–100 times more toxic than Cr(III) to

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141

microorganisms, most of the reports on the impact of Cr on biological activities in soils has been concerned with Cr(III). In these studies, not much attention has been given to the speciation of Cr after its application as Cr(III). In longterm contaminated soils, as at the Mount Barker site in South Australia, Cr(VI) is invariably present at levels toxic to microorganisms.

V. Remediation of Cr-Contaminated Water and Soils Remediation of soils, water, and sediments contaminated with metal or organic pollutants has been studied extensively in the past two to three decades. Processes developed for remediation of environments contaminated with chrome wastes are more suited for aquatic systems than for terrestrial systems. Traditional methods, used especially for wastewaters, involve chemical or electrochemical reduction of Cr(VI) to Cr(III), precipitation of the latter, and its removal by filtration or sedimentation (Eary and Rai 1988). Chemical methods are generally not cost-effective and may themselves generate hazardous byproducts (Fendorf and Li 1996). Microorganisms are capable of altering the redox state of Cr by reducing Cr(VI) to Cr(III) through direct (enzymatic) or indirect (via iron reduction, sulfate/sulfur reduction, or sulfur oxidation) processes. A. Remediation Technologies for Wastewater and Solutions Biosorption Sequestration and immobilization of heavy metals, especially in the solutions of effluents and wastewater, can be accomplished through biosorption, a passive process of metal uptake, using dead biomass in particular (Gadd 2000). Biosorption is essentially a nondirected physicochemical complexation reaction between dissolved metal species and charged cellular components that involves sorption or complexing of metals to living or dead cells. The precipitation or crystallization of metals leading to their sequestration can take place at or near the cell. Also, insoluble metal species can be physically entrapped in the microbially produced extracellular matrix or precipitated in bacterial or algal exudates (Volesky and Holan 1995). Extracellular matrices may consist of neutral polysaccharides, uronic acids, hexosamines, and organically bound phosphates that are capable of complexing metal ions. Metabolically mediated accumulation is usually intracellular and linked to the control of plasmid linked genes (Shumate and Strandberg 1985). Yeasts and bacteria as well as algae can effectively sequester metals in solutions (Kratochvil and Volesky 1998) because of their metal-binding capabilities. Algae such as Scenedesmus, Selenastrum, and Chlorella are known to bioaccumulate metals (Brady et al. 1994). The functional groups present in the cells and cell walls of fungi and algae can serve as the probable sites for biosorption of metals. For instance, the amino group of chitin (R2-NH) in the alga Sargassum and chitosan (R-NH2) in fungi are probably the effective binding site for Cr(VI) (see Kratochvil and Volesky 1998). Functional groups such as chitin and

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chitosan, however, seem to contribute only 10% of the metals sequestered by the biomass. Biosorption, using especially dead biomass, is a cost-effective technology for removal of heavy metals, and is as effective as ion exchange, but is yet to be exploited commercially. Biosorption research was confined to mostly cations, and there is a need for research on uptake of anions by biomass such as Cr. Biosorption of Cr(VI) is often followed by its bioreduction to less toxic Cr(III) and eventual precipitation of the latter. Bioreduction has been used for removal of Cr(VI) from wastewater systems in the Metex process (Linde AG, Germany) anaerobic sludge reactor, the Bio-Substrat process (Dr. Furst Systems and BKTBurggraf, Germany) anaerobic microcarrier reactor, and the Agarkar Research Institute chromate reduction process (Pumpel and Paknikar 2001). Biosorption is not suitable for detoxification of solid Cr wastes in soils. Biofilms in Bioreactors Bacterial biofilms have been recommended as an efficient means of remediating contaminants in the environment because biofilms provide tolerance to desiccation, a high level of pollutants, and other stress factors. Smith and Gadd (2000) used a mixed culture sulfate-reducing bacterial film for reduction of hexavalent Cr. In the presence of lactate as the carbon source and sulfate, 88% of 500 µmol Cr(VI) was removed from the solution with bacterial biofilm as insoluble Cr(III) in 6 hr. Because sulfide, a reductant of Cr(VI), was not detected in the medium and no reduction occurred in uninoculated medium, dissimilatory chemical reduction was not involved in Cr(VI) reduction. Evidently, Cr(VI) reduction in sulfate-reducing bacterial films was biologically mediated, presumably by enzymes. It is also possible to recover the insoluble or precipitated Cr(III) from the bacterial films. There is promise for using this biofilm technology for detoxification of Cr wastes in a bioreactor. Immobilized Cells Cells immobilized on polyacrylamide gel can be used for effective detoxification and removal of metals in solution from effluents in a reactor. Intact cells of the sulfate-reducing bacterium Desulfovibrio desulfuricans, immobilized on polyacrylamide gel, reduced about 80% of 0.5 M Cr(VI) with lactate or H2 as the electron donor and Cr(VI) as the electron acceptor (Tucker et al. 1998). Insoluble Cr(III) accumulated on the surface or interior of the gel. Immobilized cells also effected the reduction of other oxidized metals, Mo(VI), Se(VI), and U(VI). Immobilized cells may be useful for detoxification of Cr(VI) in bioreactors. Bioreactor Using Living Microorganisms Rajwade and Paknikar (1997) developed an efficient chromate reduction process using a strain of Pseudomonas mendocina MCM B-180 for treatment of chromate-containing wastewater. The bacterial strain used was resistant to 1600 mg Cr(VI)/L and reduced 2 mM chromate [100 mg Cr(VI)/L] in 24 hr. In 20-mL continuously stirred bioreac-

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tors containing this bacterium and sugarcane molasses as a nutrient, 25–100 mg chromate/L was removed within 8 hr (Bhide et al. 1996; Pumpel and Paknikar 2001). Efficiency of this bioremediation process is enhanced by anaerobiosis. B. Remediation Technologies for Chromium Wastes in Soils Traditional and innovative methods to manage Cr(VI)-contaminated soils have been reviewed by Higgins et al. (1997). The techniques chosen are mainly based on the feasibility and cost at that particular location and the concentration of Cr(VI) in the polluted soils. Although the total Cr concentration is important, in remediation technologies utmost consideration is given to Cr(VI) levels because of its carcinogenicity and mutagenicity. The guideline for risk-based cleanup of soil (USEPA 1996c) is 390 mg Cr kg−1 based on the ingestion pathway and 270 mg Cr(VI) kg−1 for human exposure by inhalation (USEPA 1996b). There is no comparable permissible soil level for Cr(III). The permissible limit for Cr(VI) in potable water is 0.05 mg L−1 (USEPA 1996a). The selection of the remediation technique for Cr-contaminated sites depends on the (1) size, location, and history of the site; (2) soil characteristics such as structure, texture, and pH; (3) type and chemical state of the contaminants; (4) the degree of contamination; (5) desired final land use; and (6) technical and financial means available. Advances in understanding the chemistry and toxicity of Cr compounds have led to efforts to remediate Cr-contaminated soil (James et al. 1997). Some of the important techniques used are excavation and disposal, soil washing, soil flushing, solidification (ex situ and in situ), vitrification, chemical and biological reduction, and phytoremediation; these have their own advantages and disadvantages (see Table 5 in Avudainayagam et al., this volume). Most appropriate technology is based on the concentration of Cr(VI) present in the contaminated soils, nature of the contamination, feasibility, and cost at that particular location. Of all these methods, bioremediation and phytoremediation have been most widely used because they are economical and do not release further wastes or harmful byproducts into the environment. Remediation Using Fe and Organic Amendments The main aim of current soil amendment techniques for Cr removal from soils is to promote irreversible reduction of Cr(VI) to Cr(III) and its hydroxides. Reduction of Cr(VI) can be achieved by incorporation of important reductants such as divalent iron, organic matter, and organic acids (James 1996). The Cr(VI) reduction reactions are as follows: Reduction with Fe and Fe compounds Fe + CrO42− + 0.5 H2O 6 Fe + 2 CrO + 13 H2O 2+

2− 4

> Fe(OH)3 + 0.5 Cr2O3 > 6 Fe(OH)3 + Cr2O3 + 8 H+

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Reduction by organic compounds (e.g., hydroquinone) 1.6 C6H6O2 + CrO42− + 2 H+

> 0.5 Cr2O3 + 1.5 C6H4O2 + 2.5 H2O

Irreversible reduction of Cr(VI) by Fe(II) to insoluble Fe-Cr(III) hydroxides is used as the major remediation strategy for chromate-contaminated soils. Amendments with Fe-bearing minerals along with organics could be effectively used for reduction of Cr(VI) and precipitation to Cr(III) complexes. Further details on remediation of soil Cr wastes using Fe and organic amendments are given by Avudainayagam et al. (this volume). Organics, as biosolids or other sludge materials, provide a diverse inoculum of microbes that can enhance Cr(VI) reduction. Conversely, organic sources have been used for Cr(VI) reduction extensively as an amendment to aid reduction processes (James and Bartlett 1983a; Buerge and Hug 1998). Losi et al. (1994a) applied 0, 12, or 50 t ha−1 of cow manure to soil irrigated with Crcontaminated groundwater, with and without alfalfa plants, to effectively reduce the Cr(VI) in the soil and reduce its transport through the irrigation water. Cr(VI) reduction (51%–98%) increased with an increase in organic matter loadings and contact time with the organic matter. More than 90% of the Cr was rendered immobile and less than 0.5% was taken up by alfalfa, minimizing the transport of Cr(VI) to drainage water. Organic amendments are known to enhance the reduction of Cr in soils by indigenous microflora (Cifuentes et al. 1996), directly involving enzymatic action or indirectly via iron and sulfur redox systems (biotic–abiotic coupling). In soils rich in dissolved organic carbon, formation of soluble Cr(III) complexes may be prone to reoxidation to Cr(VI) (Buerge and Hug 1998). There is evidence to suggest that aromatic contaminants such as phenol, 2chlorophenol, and p-cresol are suitable electron donors for Cr(VI) reduction (Shen et al. 1996). Chromium-reducing microbes may then be able to simultaneously remediate organic contaminants as well. The success of bioremediation processes mainly depends on the level of Cr contamination, the Cr(VI)-reducing efficiency of microorganisms, the stability of Cr(III) complexes formed, and conditions that are not conducive for the formation and occurrence of Mn oxides. C. Bioremediation Bioremediation has been used as a strategy employing introduced or indigenous microorganisms for complete transformation of organic pesticides to harmless end products such as CO2 and H2O. Likewise, microorganisms can transform inorganic pollutants, not necessarily completely, but to compounds with decreased solubility, mobility, and toxicity. For instance, as stated in Table 1, microorganisms can transform toxic and reactive Cr(VI) to less toxic Cr(III). Cr(VI) Bioremediation Technology A wide range of microorganisms exhibits an exceptional capacity to detoxify Cr(VI) by converting it to less soluble and

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much less toxic Cr(III) (see Table 1). This capacity is harnessed in bioremediation technology for Cr(VI) wherein the microbial strains are multiplied to a desired population level and pumped into soil or sediments in reactors to promote Cr reduction. The bioremediation efficiency can be enhanced by supplements with organic matter and other nutrients in the water or soil to promote the growth of the introduced microorganisms. The addition of organic sources to the soil can promote the proliferation of indigenous Cr(VI)-reducing microorganisms as well because Cr(VI) reducers, both aerobic and anaerobic, are ubiquitous in the soil environment. Losi et al. (1994b) decontaminated large volumes of Cr(VI)-contaminated water by passing it through an organic amended (cattle manure) soil. Indigenous soil microorganisms augmented by the organic amendments were largely involved in the reduction of Cr(VI) in the water, followed by precipitation and immobilization of the Cr(III) formed. In in situ techniques, nutrients are pumped along with aeration to promote the Cr reduction by aerobic Cr(VI)-reducing bacteria. Some Cr-reducing bacteria and algae have been efficiently used in the treatment of Cr-rich wastewater (Fude et al. 1994; Losi et al. 1994c; Cifuentes et al. 1996; see also Section V.A). Bioreactors are cost-effective and are effective for decontamination of Cr(VI)-contaminated wastewater. However, success has been limited for large-scale decontamination of Cr(VI)-polluted complex soils. For treatment of soils enriched with chromite ore processing residue, a technique involving the use of organic-rich acidic manure along with chrome-reducing microbes to effectively reduce the Cr(VI) in the waste has been developed (Fig. 3). This layer is positioned below the Cr-rich waste, and Cr(VI) leaching from the waste is effectively reduced in the organic layer, thereby preventing further contamination of groundwater (James 1996; Higgins et al. 1997). As described by Losi et al. (1994c), bioremediation of Cr(VI)-contaminated soil is achieved by either direct or indirect biological reduction. Most of the direct microbial reduction would be expected on surface soils. In the subsurface layers, indirect biological reduction of Cr(VI) involving H2S can be predominant and very effective, especially in situations where in situ stimulation of sulfatereducing bacteria is achieved through the addition of sulfate and nutrients. H2S, diffused into inaccessible soil pores, promotes the reduction of not only Cr(VI) but also Mn oxides involved in reoxidation of Cr(III). This method has shown some promise for remediation of Cr(VI)-contaminated soils when applied to an anaerobic bioreactor system (Losi et al. 1994c). Anaerobic Packed-Bed Bioreactor Anaerobic Cr(VI)-reducing microorganisms are known to be ubiquitous in soils (Turick et al. 1996). Anaerobic chromate-reducing strains have been successfully used for the reduction and sedimentation of tannery wastes (Smillie and Loutit 1982; Turick et al. 1996; Schmieman et al. 1997). Turick and his group have developed an anaerobic bioprocess for Cr(VI) reduction using a mixed culture of soil isolates or indigenous microorganisms

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Fig. 3. Bioremediation of chromite ore processing residue in soil using organics and microorganisms (MO).

in a packed-bed bioreactor containing ceramic packing or DuPont Bio-Sep beads (Turick et al. 1997, 1998). There is evidence to suggest that organic contaminants such as aromatic compounds are suitable electron donors for Cr(VI) reduction (Shen et al. 1996). Consequently, chromium-reducing microbes may then be able to simultaneously remediate organic contaminants as well. Outlook for Engineered Microorganisms Cr(VI) reduction by a wide range of microorganisms is of environmental and biotechnological significance. Bioremediation of chromate-polluted environments often poses two major problems: (1) inability of introduced Cr(VI)-reducing microorganisms to establish and function at sites polluted with mixtures of contaminants, and (2) biodegradation rates not adequate to achieve acceptable residue levels within an acceptable time frame. Several strategies have been proposed to enhance the rates of bioremediation of pollutants in such inhospitable environments. One of the approaches is to develop novel engineered strains with increased Cr(VI)-reducing efficiency for such situations. Gonzalez (2002) cloned two bacterial genes encoding different soluble chromate reductases (class I and class II) that reduce Cr(VI) to Cr(III). Each class has several close structural homologues in other bacteria. Five of these proteins, overproduced in pure form, could reduce chromate and quinones. Class II proteins could also reduce nitroaromatic compounds. Efforts are underway to use these genes and proteins directly in bioremediation of chromate-polluted environments.

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Natural Attenuation Natural attenuation involves in situ physical, chemical, and biological processes to decrease the concentration of a contaminant in the environment over time without human intervention (National Research Council 2000; Suthersan 2002). Biotransformation plays a major role in the natural attenuation of several contaminants in long-term contaminated sites. In a long-term tannery waste-contaminated site at the Mount Barker site in South Australia, industrial discharges of the waste ceased about 25 years ago. Analysis of samples revealed almost the same Cr(VI) levels in the soil (around 40 mg kg−1) and water (up to 2 mg L−1) at 20 yr (Naidu et al. 2000b; samples collected in 1997) and 25 yr after the last waste input (Kamaludeen 2002). Thus, during 5 years (1997–2002), there was no appreciable natural attenuation of Cr(VI) at this site although the soil was rich in organic carbon (9.8%–15.7%) and harbored Cr(VI)-reducing microorganisms (Megharaj et al. 2003). Incubation of this contaminated soil without and with added cow manure under saturated conditions led to complete disappearance of Cr(VI) within 20 d, but Cr(VI) reappeared, probably because of reoxidation of Cr(III) when the saturated soil was subsequently subjected to drying. However, no decrease in the concentration of Cr(VI) occurred in the Mount Barker soil held at 70% water-holding capacity even in the presence of cow manure. Although Cr(VI) can be reduced by a wide range of aerobic microorganisms (see Table 1), its reduction in the contaminated soil occurred under saturated conditions and not at 70% water-holding capacity. Reoxidation of Cr(III) and moisture stress conditions probably explains the lack of natural attenuation of Cr(VI) in the contaminated soil at the Mount Barker site. D. Applicability of Phytostabilization to Cr-Contaminated Soil Given the literature available on phytostabilization of metals, it was evident that no attempt has been made to stabilize Cr in soils. Theoretically, exudation of organic compounds by plant roots should stimulate the microbial reduction of Cr(VI) because Cr(VI)-reducing microorganisms are known to use a variety of organic compounds as electron donors. However, there are not many reports on the use of plants for the reduction of Cr(VI) to Cr(III). In general, several factors such as site characteristics and possible risk assessment must be assessed before implementing the appropriate technique to the field. Silene vulgaris, an excluder plant (Bini et al. 2001), effectively reduced Cr(VI) to Cr(III) and restricted the less bioavailable fractions of Cr in surface soils. In a study on the uptake and translocation of Cr(III) and Cr(VI) in rice plants, Cr(VI) reduction was attributed to the plant–microbe interactions in the rhizosphere (Mishra et al. 1997). A rhizosphere with intense microbial activity can play a significant role in aiding the phytostabilization of Cr. Chen et al. (2000) reported the enhanced reduction of Cr(VI) in a wheat rhizosphere. Likewise, some aquatic plants (MelLytle et al. (1998) and possibly rice (Mishra et al. 1997) have a great potential for in situ remediation of Cr because of their ability to reduce Cr(VI) to Cr(III).

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VI. Challenges As stated by James (1996), the complex chemistry involved in Cr transformations causes unique measurement and regulatory challenges. Remediation becomes complicated in heterogeneous wastes wherein the transformation reactions are rapid and interchanging. Although treatment technologies exist for remediation of Cr in soils and water, as discussed here in the individual sections, there are some setbacks in soil systems that need to be resolved. In a complex soil system, both biotic and abiotic processes play a significant role in determining the success of the remediation of Cr(VI). One of the major problems encountered in using Cr reduction as a remediation option is the Mnassisted reversible oxidation of Cr(III) to Cr(VI) on a shift in the soil to oxidizing conditions or by natural weathering processes. An indirect role of microorganisms in Cr(III) oxidation can be envisaged when microbially produced Mn oxides mediate the chemical oxidation of Cr(III). In this regard, it is necessary to precipitate and immobilize Cr(III) to forms not available for reoxidation in Mn-rich Cr-contaminated soils. Being both environmentally friendly and cost-effective, bioremediation and phytostabilization techniques are very attractive options for remediation of heavy metals. Bioremediation using ex situ bioreactors and in situ treatment approaches, especially for Cr-contaminated soils (Gadd 2000), has been investigated; however, few detailed reports exist on targeting Cr associated with tannery wastes, especially in the long-term disposal sites. Future research should be directed toward increasing the stability of Cr(III) formed using long-term contaminated soils. Phytostabilization techniques, with appropriate vegetation and soil amendments (organic manure, phosphate fertilizer, etc.) for immobilization of the metals, are yet to be effectively explored for sites contaminated with Cr (Ward et al. 1999). Overall, there is a need to include and understand the major biotic–abiotic mechanisms governing Cr transformation to develop effective remediation technologies for complex Cr-contaminated soils. Some of the major challenges are addressed in this review, with major emphasis on the effect of tannery waste contamination on soil microbial populations and their activities, the biotic–abiotic interactions involved in Cr oxidation, and the applicability of phytostabilization techniques for Cr(VI)-contaminated soils.

Summary Discharge of Cr waste from many industrial applications such as leather tanning, textile production, electroplating, metallurgy, and petroleum refinery has led to large-scale contamination of land and water. Generally, Cr exists in two stable states: Cr(III) and Cr(VI). Cr(III) is not very soluble and is immobilized by precipitation as hydroxides. Cr(VI) is toxic, soluble, and easily transported to water resources. Cr(VI) undergoes rapid reduction to Cr(III), in the presence of organic sources or other reducing compounds as electron donors, to become precipitated as hydroxides. Cr(VI)-reducing microorganisms are ubiquitous in soil

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and water. A wide range of microorganisms, including bacteria, yeasts, and algae, with exceptional ability to reduce Cr(VI) to Cr(III) anaerobically and/or aerobically, have been isolated from Cr-contaminated and noncontaminated soils and water. Bioremediation approaches using the Cr(VI)-reducing ability of introduced (in bioreactors) or indigenous (augmented by supplements with organic amendments) microorganisms has been more successful for remediation of Crcontaminated water than soils. Apart from enzymatic reduction, nonenzymatic reduction of Cr(VI) can also be common and widespread in the environment. For instance, biotic–abiotic coupling reactions involving the microbially formed products, H2S (the end product of sulfate reduction), Fe(II) [formed by Fe(III) reduction], and sulfite (formed during oxidation of elemental sulfur), can mediate the dissimilatory reduction of Cr(VI). Despite the dominant occurrence of enzymatic and nonenzymatic reduction of Cr(VI), natural attenuation of Cr(VI) is not taking place at a long-term contaminated site in South Australia, even 225 years after the last disposal of tannery waste. Evidence suggests that excess moisture conditions leading to saturation or flooded conditions promote the complete removal of Cr(VI) in soil samples from this contaminated site; but Cr(VI) reappears, probably because of oxidation of the Cr(III) by Mn oxides, with a subsequent shift to drying conditions in the soil. In such environments with low natural attenuation capacity resulting from reversible oxidation of Cr(III), bioeremediation of Cr(VI) can be a challenging task.

Acknowledgments This project was funded by the Remediation of Contaminated Environments Program, CSIRO Land and Water, and John Allwright Scholarship to S.P.B. Kamaludeen from the Australian Centre for International Agricultural Research, Canberra.

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Index

Abiotic coupling, manganese oxidation of Cr(III), 121 Aging effect on metal availability in soils, 1 ff. Aging, increases in metal availability, soils, 9 Aging metals, bioavailability to invertebrates, 11 Aging metals, soil availability, 2 Algae, chromium toxicity, 128 Algae, environmental metals risk monitor, 23 ff. Algae test protocols for toxicity, 26 Algal Assay Procedure Bottle Test (EPA), 23 Algal culture techniques, toxicity tests, 27 Algal metal toxicity, effects measurements, 37 Algal metal toxicity, regulatory context, 25 Algal metal toxicity testing, 23 ff. Algal metal toxicity testing, acclimation/ adaptation laboratory, 41 Algal metal toxicity testing, adsorption, 39 Algal metal toxicity testing, inoculum rate, 38 Algal metal toxicity testing, lab/field extrapolation, 42 Algal metal toxicity testing, toxicity mechanisms, 39 Algal metal toxicity testing, uptake, 39 Algal metal toxicity testing, variables affecting results, 43 Algal sensitivity, toxicants, interspecific differences, 35 Algal sensitivity, toxicants, intraspecies variability, 36 Algal species used in standard toxicity testing, 34 Algal toxicity testing, test media, 28 Alumina aging in soils, 4 Anabaena flos-aquau (alga), toxicity testing, 34

Anaerobic packed-bed bioreactor, Cr(VI) water remediation, 145 APHA, American Public Health Association, 34 ASTM, American Society for Testing and Materials, 34 Bioavailability, metals in soils, defined, 1 Biofilms, chromium remediation in wastewater, 142 Bioremediation technology, Cr(VI) in soils, 144 Biotic coupling, manganese oxidation of Cr(III), 121 Cadmium, algal toxicity (table) , 38 Cadmium fixation in soils, 5 Carcinogenicity, chromium, 59 Cation exchange capacities (CECs), soils, 4 CECs (See cation exchange capacities, 4 Chelators, effect in algal toxicity testing, 30 Chemical reactivity, chromate ion, 57 Chlamydomonas reinhardii (alga), pH effects Cu toxicity, 30 Chlorella pyrenoidosa (alga), Cu & Pb testing, 28 Chlorella vulgaris (alga), toxicity testing, 34, 37 Chromate ion, chemical reactivity, 57 Chromate-contaminated soil/water, remediation, 77 Chromate-contaminated soils, remediation technologies, 79 Chromite, only major commercial product, 95 Chromium (III), chemistry 57 Chromium (III), oxidation rate by γ-MnOOH, 102 Chromium (III), threshold permissible levels in soils, 78 Chromium (V), accumulation in Cr(VI) reduction to Cr(III), 119

165

166

Index

Chromium (VI), bioremediation technology, 144 Chromium (VI), chemistry, 57 Chromium (VI) direct reduction, microbial, 104 Chromium (VI), factors affecting soil solution, 72 Chromium (VI) reductases, bacteria, 118 Chromium (VI) reduction, list of capable microbes, 106 Chromium (VI) reduction microbes, list, 106 Chromium (VI) reduction, microbial aerobically, 106 Chromium (VI) reduction, microbial anaerobically, 106 Chromium (VI), soil adsorption, 74 Chromium (VI), threshold permissible levels in soils, 78 Chromium (VI) tolerance, microbes, 103 Chromium (VI) transformation to nontoxic Cr(III), 94 Chromium (VI)-tolerant bacteria, 125 Chromium, adsorption/desorption, soils, 70 Chromium, anthropogenic sources, 95 Chromium, carcinogenicity, 59 Chromium chemistry, soils, 53 ff. Chromium, commercial uses, 54 Chromium cycle in soil and water (diagram), 60 Chromium, effects on enzymes in soils, 138 Chromium, effects on microbial processes in soils (table), 133 Chromium, effects on nitrogen fixation in soil, 139 Chromium, effects on pure cultures, microorganisms (table), 133 Chromium, effects on reductive dechlorination in soil, 140 Chromium, effects on soil dehydrogenase activity, 132 Chromium, effects on soil microbial communities, 130 Chromium, effects on soil respiration, 139 Chromium extraction, serpentine soils, 62 Chromium extraction, sludge, 62 Chromium genotoxicity, 59

Chromium, in soil solution, 67 Chromium, in tannery waste sites, 53 ff. Chromium, industrial consumption, 96 Chromium, inhibitory effects on nitrification in soil, 139 Chromium, iron effects on oxidation, 98 Chromium, irreversible reduction, Cr(VI) to Cr(III), 78 Chromium, manganese effects on oxidation, 99 Chromium, natural sources, 95 Chromium, organically complexed solubility, 69 Chromium, oxidation of Cr(III) to Cr(VI), 71 Chromium oxidation reactions, soil, 98 Chromium oxidation, soil type effects, 100 Chromium, partitioning & mobility, 66 Chromium partitioning, sludge/soils, 63 Chromium, priority pollutant, 94 Chromium, reduction potential diagram, 56 Chromium remediation in soils, 143 Chromium remediation in wastewater, 141 Chromium salts, 54 Chromium sequential extraction scheme, Chinese soils, 63 Chromium, soil mobility, water solubility, 97 Chromium, soil organic matter effects, 97 Chromium, soil solution chemistry, 67 Chromium, solid-phase speciation, 61 Chromium, stable oxidation states, 55 Chromium, thermodynamic stability of aqueous species (diagram), 68 Chromium toxicity, algae, 129 Chromium toxicity, animals, humans, 58 Chromium toxicity, fungi, 127 Chromium toxicity, general, 58 Chromium toxicity, microorganisms, 126, 128 Chromium toxicity, mycorrhizal fungi, 127 Chromium toxicity, plants, microorganisms, 58 Chromium transformation in soil, factors governing, 97

Index Chromium transformation in soil, microbe factors, 103 Chromium, various forms, 95 Chromium, water solubility, 97 Chromium-contaminated soils, microorganism remediation, 93 ff. Chromium-contaminated soils, sources, 94 Chromium-contaminated water & soils, remediation, 141 Chromium-microorganism interactions, soil, 93 ff. Cobalt fixation in soils, 6 Copper fixation in soils, 5 Copper toxicity, earthworms, 10 Croococcus paris (alga), zinc toxicity testing, 37 Culture media, algal toxicity testing, 28 Cyclotella sp. (alga), toxicity testing, 34 Dehydrogenases, chromium effects in soil, 132 Dissolved organic carbon, effects on metal aging, soils, 5 DOC (See dissolved organic carbon), 5 DTPA (See diethylenetriaminepentaacetic acid), 7 DTPA, metal extraction, soils, 7 EDTA (See ethylenediaminetetraacetic acid), 7 EDTA, metal extraction, soils, 7 Environmental bioavailability, metals in soils, 9 Environmental metals risk testing, algae, 23 ff. Extraction methods, chromium in soil, 62 Fungi, chromium toxicity, 127 ␥-MnOOH, oxidation rate of Cr(III), 102 Hormidium rivulare, Cu toxicity testing, 29 Interspecific differences, algal toxicant sensitivity, 35 Intraspecies variability, algal toxicant sensitivity, 36

167

Invertebrates, metal bioavailability in soils, 10, 11 Iron, effects on chromium oxidation, 98 Iron fixation in soils, 6 Iron transformation in soils, 4 Irreversible reduction, Cr(VI) to Cr(III), 78 ISO, International Organization for Standardization, 34 Kd (Partition coefficients), metals in soils, 15 Lead fixation in soils, 5 Leather tanning with chromium, described, 54 Light, effect in algal toxicity testing, 33 Manganese (II) oxidation to Mn(IV), microbial, 122 Manganese (II) oxidizing microbes, list, 123 Manganese effects on chromium oxidation, 99 Manganese fixation in soils, 6 Manganese oxidation, microbial mechanism, 123 Manganese oxides formed by microbes, 124 Metal aging in soils, 1 ff. Metal aging, natural soils, 6 Metal availability in soils, aging effects, 1 ff. Metal extraction from soils, 6 Metal sorption/desorption, soils, 3 Metal-contaminated soils, risk assessment, 1 ff. Metals, environmental availability, defined, 1 Metals, environmental bioavailability, defined, 1 Metals, toxicological bioavailability, defined, 1 Microbial oxidation, Mn(II) to Mn(IV), 122 Microbial reduction, Cr(VI), 106 Microcystis aeruginosa (alga), toxicity testing, 34

168

Index

Microorganisms, aging metal soil bioavailability, 13 Microorganisms, capable of oxidizing manganese, 123 Microorganisms, capable of reducing Cr(VI), list, 106 Microorganisms, tolerance to Cr(VI), 103 Mycorrhizal fungi, chromium toxicity, 127 Natural soils, metal aging, 6 Navicula pelliculosa (alga), toxicity testing, 34 Nickle fixation in soils, 6 Nitrification, Cr(III) inhibitory effects in soil, 139 Nitszchia closterium (marine diatom), Cu & Zn testing, 28 Nitzschia sp. (alga), toxicity testing, 34 Nutrients, effect in algal toxicity testing, 31 OECD (See Organization for Economic Cooperation and Development), 8 OECD, Organization for Economic Cooperation and Development, 34 OECD standard artificial soils, 8 Oxidation, chromium (III) to (VI), 71 Partition coefficients (Kd), metals in soils, 15 pH, chromium redox reaction effects, 97, 99 pH, effect in algal toxicity testing, 28 pH, effect on chromium oxidation, 71 Plants, metal bioavailability in soils, 9 Plants, metal toxicological bioavailability in soils, 10 Priority pollutants, chromium, 94 Pseudokirchneriella subcapitata (alga), algae metal monitoring, 23, 30 Reductases, Cr(VI) to Cr(III), microbial, 118 Remediation, chromate-contaminated soil/ water, 77, 141 Remediation, Cr(VI)-contaminated water, 81

Remediation technologies, chromate-contaminated soils, 79 Remediation technologies, chromium in wastewater, 141 Risk assessment, metal-contaminated soils, 1 ff Risk assessment of metals, using algae toxicity tests, 23 ff. Scenedesmus quadricauda (alga), Cu toxicity testing, 29, 34, 37 Scenedesmus subspicatus (alga), toxicity testing, 34 Selenastrum capricornutum (alga), metal monitoring, 23, 34, 37 Soil chromium, chemistry, 59 Soils, metal extraction methods, 6 Solid-phase speciation, chromium in soils, 62, 64 Standard artificial soils, OECD, 8 Standard test protocols, algae toxicity tests (table), 26 Synedra sp. (alga), toxicity testing, 34 Tannery effluents, source of Cr(VI)-reducing microbes, 115 Tannery sludge composition, 75, 76 Tannery sludge, sequential chromium extraction, 65 Tannery waste, major chromium source, 96 Tannery waste sites, chromium problems, 53 ff. Tanning of leather with chromium, described, 54 Temperature, effect in algal toxicity testing, 33 Test media, algal toxicity testing, 28 Toxicity, chromium, 58 Toxicity, chromium to microorganisms, 126, 128 Toxicity tests, algal culture techniques, 27 Toxicological bioavailability, metals in soils, 10 Toxicological bioavailability, metals in soils, defined, 1 Zinc, algal toxicity (table), 37 Zinc fixation in soils, 5 Zinc, toxicity to earthworms, 11 Zinc, toxicity to springtails, 11