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Rev Environ Contam Toxicol 178:93–164 © Springer-Verlag 2003

Chromium–Microorganism Interactions in Soils:Remediation Implications

Sara P.B. Kamaludeen, Mallavarapu Megharaj, Albert L. Juhasz,Nabrattil Sethunathan, and Ravi Naidu

Contents

I. Introduction .......................................................................................................... 94A. Forms of Chromium ....................................................................................... 95B. Sources of Chromium in Soil ......................................................................... 95C. Chromium Transformations in Soil ................................................................ 97

II. Physicochemical Factors Governing Chromium Transformations in Soil ........ 97A. Soil Physical Factors ...................................................................................... 97B. Soil pH ............................................................................................................ 97C. Organic Matter ................................................................................................ 97D. Iron .................................................................................................................. 98E. Manganese ....................................................................................................... 99

III. Microbiological Factors Governing Chromium Transformations in Soil .......... 103A. Resistance or Tolerance to Cr(VI) ................................................................. 103B. Direct Cr(VI) Reduction ................................................................................. 104C. Indirect Reduction ........................................................................................... 119D. Biotic–Abiotic Coupling in Mn Oxide-Mediated Oxidation

of Chromium(III) ............................................................................................ 121IV. Implications of Chromium Transformations on Microorganisms

and their Activities .............................................................................................. 126A. Microorganisms .............................................................................................. 126B. Effects on Soil Microbial Community ........................................................... 130C. Effect on Soil Microbial Processes and Activities ........................................ 132

V. Remediation of Chromium-Contaminated Water and Soils ............................... 141A. Remediation Technologies for Wastewater and Solutions ............................ 141B. Remediation Technologies for Chromium Wastes in Soils .......................... 143C. Bioremediation ................................................................................................ 144D. Applicability of Phytostabilization to Cr-Contaminated Soil ........................ 147

VI. Challenges ............................................................................................................ 148Summary .................................................................................................................... 148

Communicated by G.W. Ware.

S.P.B. KamaludeenThe University of Adelaide, Department of Soil and Water, Waite Campus, Glen Osmond, SA 5064,Australia andTamil 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 andRemediation, University of South Australia, Mawson Lakes Campus, Mawson Lakes, SA 5095,Australia.

93

94 S.P.B. Kamaludeen et al.

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

I. Introduction

The increasing urbanization and human population worldwide has generated anever-increasing amount of inorganic and organic wastes of domestic and indus-trial origin. Such wastes have generally been disposed onto land for centuries,relying on the soil’s capacity to decontaminate waste materials by biologicaland physicochemical means and render them harmless by adsorption or precipi-tation of potential pollutants in the wastes (Martin and Parkin 1985). Continuedand excessive loading of such wastes beyond the soil’s capacity as a sink, how-ever, has led to disastrous consequences to soil and water resources worldwide.

Industrial disposal of tannery wastes onto soil had been a common practicebefore enactment of stringent regulations in many countries including Australia.One of the major problems encountered with disposal of tannery wastes is thepresence of chromium (Cr), a heavy metal, in the waste. Chromium is widelyused in the metallurgic, refractory, chemical, and tannery industries. Chromeplating, the deposition of metallic Cr, imparts a refractory nature to materials,rendering them resistant to microbial attack and flexible over extended periodsof time (Barnhart 1997). More than 170,000 tonnes of Cr waste are dischargedto the environment annually as a consequence of industrial and manufacturingactivities (Gadd and White 1993). Of the total Cr used in the processing ofleather, 40% is retained in the sludge. Generally, tannery wastes contain Cr(III)as the predominant species of Cr together with high concentrations of organiccarbon derived from the animal hides. Because of the thermodynamic stabilityof relatively low toxic Cr(III), disposal of tannery sludge onto land and intowater 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 coun-tries, primarily because the soluble Cr species, Cr(VI), is a respiratory carcino-gen 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 groundwaterquality because it is more readily transported. For these reasons, regulatory au-thorities monitoring the contaminated sites have placed considerable emphasison remediation and rehabilitation of Cr-polluted soils. The techniques for reme-diation of Cr-contaminated soils are based mostly on transforming the toxicCr(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 maylater revert to toxic species because of the changes in soil properties, differentfarming techniques, or climatic variables. Such rapid reversible transformationsof Cr in soil have further complicated the task of determining whether Cr-bear-ing waste or waste-contaminated soil is hazardous as recorded in the Federal

Chromium–Microorganism Interactions 95

Register in 1991 (James et al. 1997) and of setting the guidelines for remedia-tion.

This review highlights (i) Cr transformations in soil, (ii) abiotic and bioticfactors governing Cr transformations in soil, (iii) effect of Cr on soil microbialactivities, and (iv) remediation of Cr-contaminated soils with special emphasison bioremediation techniques.

A. Forms of Chromium

Chromium, categorized as a heavy metal, is the 24th element in the periodictable and is prevalent in nature as the 17th most abundant element on earth.Chromite (FeOCr2O3) is the only major commercial product. Chromium occursin oxidation states ranging from Cr(II) to Cr(VI) (Avudainayagam 2002; seealso Avudainayagam et al., this volume), but Cr(III) and Cr(VI) are the twostable oxidation states of Cr in soil and water environments.

Cr(VI), the form used in many industrial applications and also formed inthe environment through oxidation of Cr(III), is relatively more water soluble,bioavailable, reactive, and toxic than Cr(III). Chromium compounds are muta-genic (Venitt and Levy 1974) and teratogenic (Bauthio 1992). Generally, Cr(III)has a low toxicity, whereas Cr(VI), inhaled through dust, is recognized by theInternational Agency for Research on Cancer and by the U.S. Toxicology Pro-gram as a pulmonary carcinogen (Barceloux 1999). In addition, Cr(VI) causesirritation to and corrosion of the skin and respiratory tract in humans and anallergic contact dermatitis known as eczema.

B. Sources of Chromium in Soil

Chromium was first discovered in crocoite in 1798 and named after the brightcolor of Cr compounds. In the soil environment, it is derived from both naturaland anthropogenic sources.

Natural Sources Chromium is found preferentially in ultrabasic and basicrocks, feldspar materials in particular. On average, the earth crust contains 3,700mg Cr kg−1, most of which resides in the core and mantle (Nriagu 1988). TheCr concentration of the inner core is about 12,100 mg kg−1 (Liu 1982). Most ofthe chrome ores are located in three major places: deposits of the bushved com-plex of South Africa, the great dyke in Zimbabwe, and the kemi intrusion ofFinland (DeYoung et al. 1984). The global production of Cr annually amountsto 107 tons.

Anthropogenic Sources Chromium compounds are widely used in many indus-tries, especially in metallurgical, refractory, and chemical manufacture. Theseindustries use low-grade chromite ores for numerous applications including pig-ment manufacture, metal finishing, corrosion inhibition, organic synthesis,leather tanning, and wood preservation (USEPA 1988).


The annual Cr consumption in different industries is given in Fig. 1. Theleather industry alone accounts for 40% of the worldwide Cr usage. Large-scaledisposal of tannery wastes has significantly contributed to Cr contaminationin soils and water worldwide. Most of the Cr reaches the soil by improper dis-posal of industrial wastes, spills, or faulty storage containers (USEPA 1984).Approximately 50,000 tonnes of Cr-rich solid wastes are disposed onto landannually from tannery industries alone. The long-term disposal of tannerywastes has led to extensive contamination of agricultural soil and groundwaterin several countries, including Australia, China, India, Bangladesh, Nepal, Paki-stan, Spain, and Brazil. About 50,000 hectares of land have been rendered bar-ren by this activity in India and Bangladesh alone (ACIAR 2000). A long-termcontaminated site at Mount Barker near Adelaide (South Australia) revealed thepresence 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).

Leather40%

Others10%

Refractory3%

Pigments15%

Woodpreservation

15% Metalfinishing

17%

Fig. 1. Chromium consumption in different industries.


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 theenvironment (McGrath and Cegarra 1992), and becomes more inert over timeas a result of its precipitation. Following addition of Cr-rich wastes to soil, Crundergoes rapid transformations and attains a dynamic equilibrium betweenCr(III) ⇔ Cr(VI) through a combination of physical, chemical, and biologicalprocesses. The major processes governing Cr transformations in soils includeadsorption or desorption, redox conditions, and precipitation or dissolution (Nie-boer and Jusys 1988).

II. Physicochemical Factors Governing ChromiumTransformations in Soil

Transformations of Cr in complex and dynamic soil systems are governed byseveral physical, chemical, and biological factors.

A. Soil Physical Factors

The influence of soil physical factors in relation to Cr transformations is notdocumented in detail, except for a few reports. The adsorption and desorptionprocesses are governed largely by the bulk and particle density of the soils andthe type and amount of clay and organic matter content. In peat soils, Cr(III) isthe more prevalent species because of its binding to organic and mineral frac-tions. However, in sandy and clay soils, Cr(III) tends to oxidize because of itssolubility in easily extractable fractions of soil solution (Milacic and Stupar1995).

Chromium behaves both as anion and cation depending on its speciation. Thechemical 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). LowpH 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 ofCr(VI) to Cr(III). Organic matter during its oxidation reduces Cr(VI) (Bartlettand Kimble 1976). Citric and fulvic acids and water-soluble extracts of air-dried


Direct oxidation Indirect oxidation

Cr

oxidisers

?

Cr III

Cr VI

Mn(II)/Fe(II)

Mn(IV)/Fe(III)

Mn/Fe cycle

Fig. 2. Possible direct and indirect Cr oxidation reactions in soil.

soils form soluble complexes with Cr(III) (James and Bartlett 1983a). Organi-cally complexed Cr(III) may remain soluble whereas metal ions are quickly ad-sorbed and precipitated. Complexed Cr(III) hydroxides differ in their solubility.Freshly precipitated Cr(III) such as CrCl3 or Cr(OH)3 is highly soluble comparedto aged precipitates and hence becomes amenable to oxidation. Cr oxidationoccurs in the following order: freshly precipitated Cr(OH)3 > Cr citrate > agedCr(OH)3 in citrate > aged Cr(OH)3 (James and Bartlett 1983b). The solubility ofthese forms of Cr determines their access to microorganisms and the extent ofCr oxidation. In soils that are low in organic matter incorporation of organic-rich wastes has been recommended to promote the reduction of Cr(VI) (Bartlettand Kimble 1976).

D. Iron

Chromium transformations are influenced by other ionic species released duringchanges in pH. For example, at low pH, the presence of ferrous [Fe(II)] ironincreased the rate of Cr reduction (Palmer and Wittbrodt 1991). The reductionof Cr(VI) to Cr(III) in the aqueous (aq) phase was rapid, as described by thefollowing reaction:.

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


At pH values greater than 4.0, brown precipitates were observed (Pettine etal. 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, presumablyCr0.25 Fe0.75(OH)3, was not conclusively identified (Patterson et al. 1997). At lowpH, the addition of small amounts of iron alone can increase the rate of Crreduction.

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 proposedto explain the chemical reduction of Cr by Fe. First, humic substances convertFe(II) to Fe(III) which in turn reduces Cr. Second, Fe(II) can form FeCrO4,which complexes with chromate. Amorphous iron sulfide minerals such asmackinawite (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). Partiallyoxidized iron was also effective and widely used for subsurface remediation. Insubsoils, a combination of Fe(II), organic matter and low pH (4.2–4.3) governthe reduction of Cr(VI) (Powell et al. 1995).

E. Manganese

Manganese oxides have been implicated in the chemical oxidation of Cr(III) inthe soil environment (Kim et al. 2002). Bartlett and James (1979) were the firstto report Mn-mediated oxidation of Cr in soils with a pH above 5.0, providedthe soil was fresh and moist. The amount of Cr oxidized was proportional tothe amount of Mn reduced (exchangeable) and also to the amount of Mn reduc-ible by hydroquinone. The minimum amount of MnO2 necessary for completeoxidation of Cr in soil is not known. Knowledge of the energetics of the reactionin relation to the availability of oxidizable Cr(III) would be desirable.

For optimum oxidation of Cr(III) by manganese oxides, the surface of thelatter must be relatively free of specifically adsorbed Mn(II) and other heavymetals. The Cr(III) approaches a receptive surface, is quickly oxidized to theanionic 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 oxidesavailable for oxidation.

pH Manganese-mediated Cr oxidation is dependent on soil pH and is normallyobserved in any soil with a pH of 5.0 and above. Up to pH 5.5, Cr(III) oxidationwas increased by naturally occurring δ-MnO2 (Fendorf and Zasoski 1992). How-ever, 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 reac-tion with synthetic β-MnO2. The extent of Cr(III) oxidation at lower pH is lim-ited by the strong adsorption of anionic Cr(VI), thereby inhibiting contact be-tween active oxidizing sites on β-MnO2 (Eary and Rai 1987).


Chromium (III) oxidation by naturally occurring δ-MnO2 was suppressed aspH 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, initialsurface area, and ionic strength (Fendorf and Zasoski 1992).

Oxidation of Cr(III) by a MnO2 preparation proceeded rapidly at pH 5.5 and7.5 with identical rates, slowly at pH 3, and very slowly at pH 1 (Amacher andBaker 1982). In addition to pH-dependent charge characteristics of oxide miner-als, Mn oxide surfaces might also have permanent negative charges as substitu-tion of Mn(II) and Mn(III) for Mn(IV) occurs during their oxidation. Becauseof the rapid redox transformations and specific adsorption continually takingplace on manganese oxide surfaces, these charges will be temporary. Exchange-able Cr(III) is not found in soils with pHs greater than 4.5 or 5. As the pH of a1 µM Cr(III) solution was lowered, amount of oxidation by a dilute soil suspen-sion ranged from 20% at pH 7.5 to 100% at pH 3.2. In some systems, the effectsof pH on charge characteristics and surface behavior of manganese oxides canmask the effects of pH on Cr speciation and solubility.

Soil Type Studies conducted in whole soils to observe the Cr oxidation haveshown that the extent of Cr(III) oxidation varied with clay content and the pro-portion of waste materials containing Cr(III) (Bartlett 1985, 1986). Landdisposal of Cr and the associated health effects have been fully reviewed byChaney et al. (1981). Behavior of Cr(VI) in organic waste is similar to slow-release nitrate fertilizers, and chromium in sludge could release low levels ofCr(VI) over a period of years. Also, the Cr associated with high molecularweight ligands is not readily oxidized (Amacher and Baker 1982). These authorsreported that sludge-borne Cr was not oxidized even over a 4-yr period; how-ever, 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, Jamesand Bartlett (1983a) observed that soils incubated under moist conditions re-leased 0–41 µmol Cr(VI) L−1 after 4 yr of incubation. In 1-m2 field plots contain-ing 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 thefour important factors of waste oxidation potential (WOP), soil oxidation poten-tial (SOP), soil reduction potential (SRP), and soil-waste pH modification value(PMV) as per the following formula: PCOS = WOP + SOP + SRP + PMV (Cha-ney et al. 1996). Higher values indicated a greater possibility for Cr oxidationin soils. This approach can be used as a quick screening tool to determine theoxidative ability of soils. Recently, the nature of Cr precipitates on goethite (acrystalline ferric oxide) and silicon dioxide was studied using scanning forcemicroscopy. This study demonstrated a new concept in that the ability of the


precipitating phase to fit the crystal lattice structure of the material on which itis precipitating can cause a significant change in the form and stability of Crprecipitates. Chromium precipitated on goethite spread evenly on the surfaceand had a low extractability with oxalate compared to Cr precipitating on silicaas clumps of Cr(OH)3 that were extracted more rapidly with oxalate (Fendorf etal. 1996) Thus, studies of the chemical nature of Cr precipitated in differentsoils might also add to our understanding of how soil chemistry can influencethe potential oxidation of Cr(III) (Chaney et al. 1996).

Mineralogy of Mn Oxides Manganese oxides have a high adsorptive capacityfor metal ions, thus potentially providing local surface environments in soil. Theoxidation of Cr(III) to Cr(VI) occurs after adsorption of Cr to the Mn, withsimultaneous 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, soilpH, 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. Therewas a proportional increase in oxidation of Cr(III) as the surface of Mn oxideincreased. 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 thedissolution of Mn oxides and enhanced the oxidation of Cr(III). As the pHincreased 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 than90% Cr(III) oxidation. The reaction was also faster (24 hr) with birnessite thanwith pyrosulite (β-MnO2). pH had a similar effect on oxidation, i.e., acidic pHfavored more oxidation. There was no difference in rate of oxidation until pH4.0, whereas beyond this pH there was a decrease up to pH 5.2. Cr oxidationwas also observed at pH 6.3, 8.3, and 10.1 in spite of the low solubility ofCr(III) at these pH levels. Increases in Cr(III) concentration (200–800 µM) re-tarded 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 forevery mole of Cr(VI) formed. Generally, oxidation of Cr(III) decreased withincreasing pH and Cr(III) concentration (Eary and Rai 1987), and severalhypotheses have been postulated to explain the inhibition.


During the Cr(III) and MnO2 reaction, only a portion of MnO2 is availablefor 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 poison-ing 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 ofMn(II) the negative charge on MnO2 surfaces is lowered; this in turn increasesthe pH of the surface because of −OH ions, and Mn(II) becomes autoxidized byatmospheric oxygen. When Mn(II) becomes autoxidized, the surface becomesmore negative and is available for further Cr(III) oxidation. However, Fendorfet al. (1993) showed that oxidation of Cr(III) was not inhibited by the additionof Mn(II) to the system.

In contrast, Eary and Rai (1987) proposed that Cr(VI) formed was a limitingfactor in Cr(III) oxidation. The initial Cr(III) oxidation was instantaneous, andCr(VI) formed becomes strongly bound to β-MnO2 surfaces, especially at acidicpHs. Moreover, this adsorption also retarded the dissolution of MnO2. The disso-lution 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. Thedifference in mechanism between δ- and β-MnO2 was attributed mainly to thedifferences 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 alsolead 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 anintermediate (Johnson and Xyla 1991) The oxidation kinetics of Cr(III) toCr(VI) on the surface of manganite (γ-MnOOH) is a function of Cr(III) andmanganite concentration, pH, ionic strength, and temperature. The reaction isfirst 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 thisconcentration, the reaction is inhibited. The reaction is independent of pH andionic strength.

Considering the chemical speciation, the overall reaction at pH 4.5 can bewritten 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 withother Mn oxides. Such fast rates of Cr(III) oxidation by γ-MnOOH contributesignificantly in the cycling of Cr in natural water systems.

Nakayama et al. (1981) found that in seawater only 10% of a 10−5 M solutionof Cr(III) was oxidized with 30 mg γ-MnOOH L−1 in 100 hr. The low oxidationnote may be linked to the organic substances in natural waters. Experimentswith salicylate clearly indicated the inhibition of Cr(III) oxidation reaction withγ-MnOOH by organic ligands.


III. Microbiological Factors Governing ChromiumTransformations in Soil

Biological factors also play a major role in the transformations (Cr reductionin particular) and mobilization of Cr in soils. A wide variety of heterotrophicmicroorganisms is involved in the reduction of Cr(VI) to Cr(III), aerobically oranaerobically depending on the organism, in both soil and water environments(Lovley and Phillips 1994). Evidence suggests that Cr(VI)-reducing microorgan-isms are ubiquitous in soils and can enhance the detoxification of Cr(VI) underideal physicochemical conditions (Turick et al. 1996). Chromium(VI)-tolerantand -sensitive bacteria, with ability to transform Cr(VI) to Cr(III), occur widelyin diverse ecological conditions: water, sediments, and soil (Losi et al. 1994a).The important microbial processes influencing Cr transformations are consid-ered in this section.

A. Resistance or Tolerance to Cr(VI)

A variety of mechanisms have been implicated in the adaptation, tolerance, andresistance 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)] toa less toxic form. Chromate tolerance in microorganisms has probably evolvedduring their long-term exposure to naturally occurring or anthropogenic sourcesof 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 harborbacterial populations resistant to Cr(VI). Thus, Bader et al. (1999) found thatboth contaminated soil with a high Cr level of 12,400 mg/kg soil and two uncon-taminated soils harbored aerobic bacterial populations resistant to Cr(VI). How-ever, bacterial populations resistant to Cr(VI) at concentrations as high as 500µg/mL could be isolated only from the uncontaminated soils and not from thecontaminated soil samples. In contrast, fungal colonies resistant to Cr(VI) atconcentrations as high as 1000 µg/mL were routinely isolated from both uncon-taminated and contaminated soils. Evidently populations resistant to Cr(VI) haveevolved in soils, not necessarily related to their previous exposure to Cr. Con-taminated soil with a high Cr content of 12,400 mg/kg soil contained much asmaller and less diverse microbial population than that in the uncontaminatedsoils. For effective bioremediation, Cr(VI) resistance through decreased uptakealone may not be desirable; decreased uptake coupled with its ability to reduceCr(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 chromo-somal mutations (Cervantes and Silver 1992). Chromosomal and plasmid deter-minants operate by different mechanisms, as evident by their additive effects.Mutations in bacterial cells caused by DNA damage by Cr(VI) have been re-ported. Plasmids conferring chromate resistance have been reported in Strepto-


myces lactis (Efstathiou and McKay 1977), Pseudomonas aeruginosa (Summersand 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. Accu-mulation of chromate in the resistant strain was 2.2 times less than that in aplasmid-cured sensitive strain. Chromate-resistant strains of P. fluorescens(Ohtake et al. 1987) and A. eutrophus (Nies et al. 1989), however, transportedsulfate to the same extent as the plasmid-cured chromate-sensitive strains. Chro-mate resistance in P. fluorescens was not related to sulfate transport.

Currently, eight proteins of the chromate resistance (Chr) family have beenfully sequenced (Nies et al. 1998). Among these, Chr proteins from Pseudomo-nas aeruginosa (Chr Pae) (Cervantes et al. 1990) and Alcaligenes eutrophus(Chr Aeu) have been functionally characterized for their role in chromate resis-tance. Chr proteins conferred decreased uptake of chromate in chromate-resis-tant P. aeruginosa (Cervantes et al. 1990) and A. eutrophus (Nies et al. 1990).Recent evidence suggests that decreased accumulation of chromate in P. aerugi-nosa, conferred by chromate-resistant protein (ChrA), was associated with itsactive efflux from the cytoplasm driven by the membrane potential (Alvarezet al. 1999). Everted membrane vesicles of P. aeruginosa harboring the Chr Aplasmid accumulated four times more chromate than did the vesicles from plas-mid-cured cells. There is no evidence yet, however, for chromate efflux as amechanism for reduced accumulation of Cr in chromate-resistant A. eutrophus.

Alcaligenes strain CH34 has determinants encoding inducible resistance tochromate whereas A. alcaligenes AE104 (plasmid-free derivative of CH34) ischromate sensitive. A lux-coupled chromate biosensor, A. eutrophus AE104(pEBZ141), carrying chrulux transcriptional fusion, developed using a clonedpart of plasmid pMOL28 carrying a chromate-resistant determinant from strainCH34, was specific for chromate (Peitzsch et al. 1998). Data from this study oninteractions between chromate resistance and chromate reduction and betweensulfate concentration and chromate induction showed that chromate resistancewas best induced by chromate and bichromate whereas induction by Cr(III) was10 times lower. Sulfate starvation increased uptake and reduction of chromatein strain 104.

B. Direct Cr(VI) Reduction

In soils, microbial Cr reduction may occur directly or indirectly. In the directmode, Cr is taken up by the microbes and then enzymatically reduced (Komoriet al. 1990b; Losi et al. 1994c; Lovley and Coates 1997), while in the indirectmode, products (reduction or oxidation) of microbial decomposition in the soilsuch as H2S mediate the reduction of Cr(VI) (DeFilippi and Lupton 1992). Di-rect microbial reduction of Cr(VI) was first reported in the 1970s (Lebedevaand Lyalikova 1979; Romanenko and Korenkov 1977) when certain Pseudomo-nas strains, isolated from chromate-containing sewage sludges, could reduce


chromate, dichromate, and crocoite during anaerobic growth. Since then, severalbacteria with exceptional ability to reduce Cr(VI) have been isolated from Cr-contaminated and uncontaminated soil samples. Microorganisms implicated indirect or indirect reduction of Cr(VI) are listed in Table 1.

Cr(VI) Reduction in Microbial Cultures Since the first reports of isolation offacultative anaerobic Cr(VI)-reducing bacteria in the mid-1970s (Romanenkoand Korenkov 1977), the literature is abundant with instances of the reductionof Cr(VI) by several microorganisms, bacteria in particular (Ohtake and Silver1995), mostly isolated from Cr-impacted environments (Table 1). Strains ofOscillatoria, Chlorella, and Zoogloea have also been reported to enzymaticallyreduce Cr(VI) (Losi et al. 1994b). But, as noticed with bacterial resistance toCr(VI), Cr(VI)-reducing bacteria have been isolated also from environmentswith minimum or no impact of Cr (Wang et al. 1989; Turick et al. 1996; Badaret al. 2000). It is also interesting to note that pure cultures of microorganismsnot previously exposed to Cr(VI) were capable of reducing it (Gvozdyak et al.1986). Although the exact mechanism is not known, microorganisms capable ofreducing Cr(VI) acquired the enzymes for degrading related compounds presentin the environment or produce the reductants that in turn reduce Cr(VI) bychemical redox reactions. Anaerobic chromate-reducing strains are prevalent insubsurface soils and probably enhance Cr reduction in this environment (Turicket al. 1996).

Chromium(VI) resistance and reduction are not necessarily interlinked. Chro-mium(VI) may be reduced by both Cr(VI)-resistant and Cr(VI)-sensitive strainsof bacteria, and not necessarily all Cr-resistant bacteria can reduce Cr(VI). Forinstance, 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 condi-tions may not be a resistance mechanism in bacteria, but a trivial side activityof 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 im-portant to use Cr(VI)-resistant microbes, with ability to reduce it. Two strainsof Pseudomonas fluorescens, one resistant and the other sensitive to Cr(VI),reduced Cr(VI) at comparable rates (Ohtake et al. 1987; Bopp and Ehrlich1988). Likewise, three Cr(VI)-sensitive bacteria from an uncontaminated soiland three Cr(VI)-resistant bacteria from two metal-stressed foundry soils and atannery readily reduced Cr(VI) anaerobically (Badar et al. 2000). Interestingly,a Cr(VI)-sensitive Bacillus sp. from the uncontaminated soil was the most effec-tive in reducing Cr(VI) among the three Cr(VI)-resistant bacterial strains frommetal-stressed soils and three Cr(VI)-sensitive bacterial strains from the uncon-taminated soil. These bacteria grew aerobically in acetate minimal medium sup-plemented with sodium chromate, but reduced Cr(VI) only anaerobically in thesuspension of resting cells of aerobically grown bacteria. Anaerobic growth ofthe bacterium at the expense of Cr(VI) as electron acceptor was negligible.Conversely, an Arthrobacter sp., isolated from a long-term tannery waste-contaminated soil, was resistant to Cr(VI) at 100 µg/mL but could not reduce it

106 S.P.B. Kamaludeen et al.T

able

1.M

icro

orga

nism

sca

pabl

eof

redu

cing

Cr(

VI)

.

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Con

sort

ium

ofsu

lfat

e-M

etal

refi

ning

was

te-

2500

mg

Cr(

VI)

/L80

%–9

5%fr

om50

–In

dire

ct;

invo

lvin

gH

2SFu

deet

al.

1994

redu

cing

bact

eria

wat

ers

2000

ppm

(SR

B)

Ent

erob

acte

rcl

oaca

eA

ctiv

ated

slud

ge52

0m

g/L

90%

rem

oval

Mem

bran

eas

soci

ated

,W

ang

etal

.19

89;

Ko-

HO

1m

ori

etal

.19

90a,

bC

rO2− 4

aste

rmin

alel

ectr

onac

cept

or

Ent

erob

acte

rcl

oaca

eA

ctiv

ated

slud

ge52

0m

g/L

90%

rem

oval

Ana

erob

icre

duct

ion

ofK

omor

iet

al.

1989

;H

O1

Cr(

VI)

whi

legr

ow-

Oht

ake

etal

.19

90;

ing

onac

etat

e,m

a-R

ege

etal

.19

97la

te,

succ

inat

e,et

ha-

nol,

and

glyc

erol

,bu

tin

hibi

ted

bygl

ucos

e

Ent

erob

acte

rcl

oaca

eA

ctiv

ated

slud

ge52

0m

g/L

Red

uces

Cr(

VI)

anae

ro-

Wan

get

al.

1990

,H

O1

bica

llyw

ithac

etat

e19

91;

Fujii

etal

.an

dca

sam

ino

acid

s19

90as

elec

tron

dono

rs;

O2

inhi

bits

Cr(

VI)

re-

duct

ion;

activ

ityas

-so

ciat

edw

ithm

em-

bran

efr

actio

nin

volv

ing

NA

DH

asel

ectr

ondo

nor

cyto

-ch

rom

c 548

Chromium–Microorganism Interactions 107T

able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Aer

omon

asde

chro

mat

ica

Cr(

VI)

redu

ced

inth

eK

vasn

ikov

etal

.19

85,

pres

ence

ofpr

opyl

1986

,19

87al

coho

lsan

dac

etat

e;ce

rtai

nhe

avy

met

als

prom

oted

orin

hib-

ited

Cr(

VI)

redu

c-tio

n/an

aero

bic

Ach

rom

obac

ter

eury

dice

,A

ceta

te,

gluc

ose/

anae

r-G

vozd

yak

etal

.19

86B

acil

lus

cere

us,

B.

obic

subt

ilis

,M

icro

cocc

usro

seus

,P

seud

omon

asae

rugi

nosa

Bac

illu

ssp

.G

luco

se/a

erob

icW

ang

and

Xia

o19

95

Pse

udom

onas

dech

rom

a-Pe

pton

e,gl

ucos

e,H

2/R

oman

enko

and

Kor

en-

tica

nsan

aero

bic

kov

1977

Alc

alig

enes

eutr

ophu

s—

—Su

lfat

est

arva

tion

led

—Pe

itzsc

het

al.

1998

AE

104

toin

crea

sed

chro

-m

ate

upta

kean

dch

rom

ate

redu

ctio

n

Bac

illu

ssp

.Q

C1–

2So

ilan

dw

ater

sam

-17

mg/

L10

0%w

ithin

22hr

Cr(

VI)

redu

ctio

nby

Cam

pos

etal

.19

95pl

esne

arch

ro-

cell-

free

extr

act;

aer-

miu

m-p

roce

ssin

gob

ical

ly/a

naer

obi-

fact

ory

cally

;C

r(V

I)re

duc-

tase

(sol

uble

NA

DH

depe

nden

t)


able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Bac

illu

ssp

.C

hrom

ate-

10–2

00m

g/L

Con

tinuo

us-f

low

bio-

Chi

rwa

and

Wan

gco

ntam

inat

edso

ilfi

lmre

acto

r,10

0%19

97a;

Wan

gan

dre

mov

alin

6–24

hrSh

en19

97

Agr

obac

teri

umra

dio-

Soil

26m

g/L

100%

rem

oval

with

inC

r(V

I)re

duct

ion

unde

rL

love

raet

al.

1993

bact

erE

PS-9

166

hrbo

thae

robi

can

dan

-ae

robi

cco

nditi

ons

Sacc

haro

myc

esce

revi

s-—

1.9

mg/

L10

0%re

duct

ion

—K

raut

eret

al.

1996

iae

Pse

udom

onas

fluo

resc

ens

——

—G

luco

se/a

erob

icB

opp

and

Ehr

lich

LB

300

1988

;C

hirw

aan

dW

ang

1997

b;W

ang

and

Shen

1997

;D

e-le

oan

dE

hrlic

h19

94

Pse

udom

onas

chro

mat

o-—

——

Seve

ral

orga

nic

com

-L

ebed

eva

and

Lya

li-ph

ila

poun

dsas

elec

tron

kova

1979

dono

rs/a

naer

obic

Pse

udom

onas

K-2

1A

erob

ical

lyre

duce

dSh

imad

aan

dM

atsu

-C

r(V

I)w

hile

grow

-sh

ima

1983

ing

ongl

ucos

e

Pse

udom

onas

ambi

gua

——

—N

utri

ent

brot

h/ae

robi

c;H

orits

uet

al.1

987;

Su-

G-1

Cr(

VI)

-red

ucta

sezu

kiet

al.

1992

NA

D(P

)H-d

epen

dent

Pse

udom

onas

stut

zeri

Met

al-c

onta

min

ated

52m

gch

rom

ate/

L58

%–8

8%C

rre

duc-

Enz

ymat

icN

AD

(P)H

-B

adar

etal

.20

00(t

wo

stra

ins)

foun

dry

soil

tion

anae

robi

cally

depe

nden

t,C

r-re

sist

ant


able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

An

unid

entif

ied

bac-

Tan

nery

52m

g/L

76%

Cr

redu

ctio

nan

-C

r-re

sist

ant,

grew

aero

-B

adar

etal

.20

00te

rium

aero

bica

llybi

cally

Pse

udom

onas

synx

anth

a,U

ncon

tam

inat

edso

il<5

mg/

L75

%–9

2%re

duct

ion

Cr-

sens

itive

,gre

wae

ro-

Bad

aret

al.

2000

Bac

illu

ssp

.an

dan

un-

anae

robi

cally

bica

llyid

entif

ied

gram

-pos

i-tiv

est

rain

Pse

udom

onas

puti

da,

P.

——

Cr(

VI)

redu

ctio

n,ae

ro-

Who

lece

lls,

cell

sus-

Ishi

bash

iet

al.

1990

fluo

resc

ens,

and

E.

bica

llype

nsio

n,ce

ll-fr

eeex

-co

litr

act

Pse

udom

onas

aeru

gino

sa—

——

Cr(

VI)

redu

ctas

ege

neJi

net

al.

2001

HP0

14tr

ansf

erre

dto

to-

bacc

opl

ant

cells

us-

ing

Agr

obac

teri

umtu

mef

acie

nsbi

nary

vect

orsy

stem

Pse

udom

onas

aeru

gino

saT

anni

ngef

flue

nt10

–25

mg/

mL

Cr(

VI)

redu

ctio

n,ae

ro-

Inba

tch

cultu

re,

dial

-K

hare

etal

.199

7;G

an-

A2C

hrbi

cally

ysis

reac

tor,

and

guli

and

Tri

path

yce

llsen

trap

ped

ina

1999

,20

01,

2002

biof

ilmin

abi

olog

i-ca

lro

tatin

gco

n-ta

ctor

;m

axim

umre

-du

ctio

nat

10m

g/m

LP

seud

omon

assp

.C

r(IV

)re

duct

ion,

aero

-In

cont

inuo

ussu

s-G

opal

anan

dV

eera

-bi

cally

pend

ed-g

row

thcu

l-m

ani

1994

ture

sP

seud

omon

asm

endo

cina

Coo

ling

tow

eref

flue

ntC

r(V

I)re

duct

ion

Bhi

deet

al.

1996

Stre

ptom

yces

sp.

—C

r(V

I)re

duct

ion

Das

and

Cha

ndra

1990

Esc

heri

chia

coli

Ace

tate

/ana

erob

icK

vasn

ikov

etal

.19

88


able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Esc

heri

chia

coli

AT

CC

Act

ivat

edsl

udge

16m

g/L

100%

redu

ctio

nin

12E

nzym

atic

redu

ctio

n/Sh

enan

dW

ang

1993

;33

456

hrgl

ucos

e,ac

etat

e,pr

o-W

ang

and

Shen

pion

ate/

aero

bic

and

1997

anae

robi

cC

ocul

ture

ofE

sche

rich

iaB

oth

from

activ

ated

Aw

ide

rang

eof

elec

-Sh

enan

dW

ang

1993

,co

liA

TC

C33

456

slud

getr

ondo

nors

for

1995

;W

ang

and

[Cr(

VI)

redu

cer]

and

Cr(

VI)

redu

ctio

nX

iao

1995

Pse

udom

onas

puti

daD

MP-

1(p

heno

lde

-gr

ader

)C

ocul

ture

ofE

sche

rich

iaB

oth

from

activ

ated

Con

tinuo

us-f

low

bior

e-W

ang

and

Chi

rwa

coli

AT

CC

3345

6sl

udge

acto

r19

98[C

r(V

I)re

duce

r]an

dP

seud

omon

aspu

tida

DM

P-1

(phe

nol

de-

grad

er)

Dei

noco

ccus

radi

odur

ans

——

—R

educ

esC

r(V

I),

Fred

rick

son

etal

.20

00Fe

(III

),U

(VI)

,an

dT

e(V

II)

Des

ulfo

vibr

iovu

lgar

is—

——

c 3cy

toch

rom

eas

Lov

ley

1993

;L

ovle

yA

TC

C29

579,

D.

sulf

u-C

r(V

I)re

duct

ase/

H2/

and

Phill

ips

1994

rica

nsan

aero

bic

Pan

toea

aggl

omer

ans

Surf

ace

sedi

men

ts—

—D

issi

mila

tory

redu

ctio

nFr

anci

set

al.

2000

SP1

onac

etat

ean

dot

her

elec

tron

dono

rsP

anto

eaag

glom

eran

sSu

rfac

ese

dim

ents

——

Ele

men

tal

sulf

urdi

s-O

braz

tsov

aet

al.

2002

SP1

prop

ortio

nate

dis-

sim

ilato

ryre

duct

ion

unde

ran

aero

bic

con-

ditio

ns


able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Eig

htfa

culta

tive

anae

r-T

anne

ryef

flue

nts

>400

µgch

rom

ate

mL

−173

%–9

4%re

duct

ion

Bot

han

aero

bica

llyan

dSr

inat

het

al.

2001

obes

(fiv

eis

olat

esof

anae

robi

cally

and

aero

bica

llyin

pep-

Aer

ococ

cus,

two

iso-

18%

–63%

redu

c-to

new

ater

late

sof

Mic

roco

ccus

,tio

nae

robi

cally

and

one

isol

ate

ofA

er-

omon

as)

Gra

m-p

ositi

vedi

chro

-T

anne

ryef

flue

nt80

mg

K2C

r 2O

7/mL

87%

ofth

eC

r(V

I)in

—Sh

akoo

riet

al.

1999

,m

ate-

resi

stan

tba

cte-

20m

gK

2Cr 2

O7/m

L20

00ri

um(A

TC

C70

0729

)re

duce

din

72hr

Pse

udom

onas

sp.

CR

B5

Woo

dpr

eser

vatio

n52

0m

gC

r(V

I)/L

23%

redu

ced

in24

hrR

espi

reae

robi

cally

and

McL

ean

and

Bev

er-

site

with

chro

mat

ed19

%in

the

pres

ence

anae

robi

cally

usin

ga

idge

2001

;M

cLea

nco

pper

arse

nate

ofA

s(V

),27

%in

vari

ety

ofte

rmin

alet

al.

2000

the

pres

ence

ofel

ectr

onac

cept

ors,

Cu(

II)

Fe(I

II),

Mn(

IV),

NO

− 2,N

O− 3,

SO2;

solu

ble

enzy

me,

mai

nly

cyto

-pl

asm

ic,

invo

lved

inre

duct

ion

Ana

erob

icen

rich

men

tA

quif

erse

dim

ents

39m

g/L

39m

g/L

in6

dC

r(V

I)re

duct

ion

and

Mar

shan

dM

cIne

rney

grow

thw

ere

depe

n-20

01de

nton

H2

Des

ulfo

tom

acul

umre

-C

r-co

ntam

inat

edse

di-

10m

g/L

Gro

ws

and

redu

ces

Teb

oan

dO

braz

tsov

adu

cens

men

tsC

r(V

I)as

sole

elec

-19

98tr

onac

cept

orin

the

pres

ence

ofbu

tyra

teas

the

carb

onso

urce

and

inth

eab

senc

eof

sulf

ate;

nogr

owth

inab

senc

eof

Cr(

VI)


able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Shew

anel

laon

eide

nsis

Ana

erob

iczo

neof

—C

r(V

)de

tect

eddu

ring

Red

uces

adi

vers

eM

yers

etal

.20

00(e

arlie

rde

sign

ated

asM

n-ri

chse

dim

ents

redu

ctio

nby

for-

arra

yof

com

poun

dsS.

putr

ifac

iens

MR

-1)

mat

e-de

pend

ent

unde

ran

aero

bic

con-

Cr(

VI)

redu

ctas

edi

tions

incl

udin

gM

n(IV

),Fe

(III

),fu

-m

arat

e,an

dC

r(V

I);

form

ate-

orN

AD

H-

depe

nden

tC

r(V

I)re

-du

ctas

eac

tivity

inan

aero

bica

llygr

own

cells

with

high

est

ac-

tivity

incy

topl

asm

icm

embr

ane

Shew

anel

laon

eide

nsis

Ana

erob

iczo

neof

—L

ive,

rest

ing

cells

re-

Prec

ipita

tes

encr

ustin

gD

aulto

net

al.

2002

Mn-

rich

sedi

men

tsdu

ced

80%

ofba

cter

ial

cells

con-

Cr(

VI)

in1

hr;

so-

tain

edC

r(II

I)or

dium

azid

ean

dhe

atlo

wer

oxid

atio

ntr

eatm

ent

stop

ped

stat

esre

duct

ion;

nore

duc-

tion

ince

ll-fr

eesu

-pe

rnat

ants

Shew

anel

laon

eide

nsis

Ana

erob

iczo

neof

—C

r(V

I)re

duct

ion

un-

Cr(

VI)

indu

cibl

e,as

so-

Via

maj

ala

etal

.M

n-ri

chse

dim

ents

der

fum

arat

e-an

dci

ated

with

anae

robi

c20

02a,

bni

trat

e-re

duci

nggr

owth

ofba

cter

ium

cond

ition

s


Tab

le1.

(Con

tinu

ed).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Thi

obac

illu

sfe

rrox

idan

s—

Cr(

VI)

redu

ctio

nby

Sist

iet

al.

1996

,19

98pr

oton

s(s

ulfi

te,

thio

-su

lfat

e,et

c.)

with

high

redu

cing

pow

erge

nera

ted

duri

ngox

i-da

tion

ofel

emen

tal

sulf

ur;

aero

bic

Thi

obac

illu

sfe

rrox

idan

s—

—C

r(V

I)re

duct

ion

bySi

sti

etal

.19

96,

1998

;pr

oton

s(s

ulfi

te,

thio

-Q

uiIn

tana

etal

.su

lfat

e,et

c.)

with

2001

high

redu

cing

pow

erge

nera

ted

duri

ngox

i-da

tion

ofel

emen

tal

sulf

ur;

rela

ted

toth

eco

ncen

trat

ion

ofpr

o-to

nsge

nera

ted;

oc-

curs

atpH

2–8,

mor

epr

onou

nced

atlo

wer

pH;

aero

bic

and

anae

robi

c,bu

tm

ore

pron

ounc

edae

robi

cally


able

1.(C

onti

nued

).

Cr

tole

ranc

eC

r-re

duci

ngId

entif

icat

ion

Sour

cele

vel

effi

cien

cyM

echa

nism

Ref

eren

ce

Bac

illu

ssu

btil

is—

——

Non

met

abol

izin

gba

cte-

Fein

etal

.20

02Sp

oros

arci

naur

eae

rial

cells

redu

ced

Shew

anel

lapu

tref

acie

nsC

r(V

I)on

cell

sur-

face

inab

senc

eof

exte

rnal

lysu

pplie

del

ectr

ondo

nors

;no

tco

uple

dto

oxid

atio

nof

bact

eria

lex

udat

es;

fast

est

unde

rac

idic

cond

ition

s

Alg

ae—

—T

rans

itory

form

atio

n—

Liu

etal

.19

95Sp

irog

yra

sp.

and

ofC

r(V

)du

ring

Mou

geot

iasp

.C

r(V

I)re

duct

ion

Osc

illa

tori

a—

—E

nzym

atic

ally

redu

ced

—L

osi

etal

.19

94b

Cr(

VI)

Chl

orel

la—

—E

nzym

atic

ally

redu

ced

—L

osi

etal

.19

94b

Cr(

VI)

Ana

baen

ava

riab

ilis

——

Chr

omat

ere

duce

din

Gar

nham

and

Gre

en18

dto

Cr(

III)

;50

%19

95of

Cr(

III)

form

edac

cum

ulat

edin

the

cells

and

50%

inth

em

ediu

m;

Cr(

VI)

redu

ctio

nas

soci

-at

edw

ithhe

tero

-cy

sts


at this concentration (Megharaj et al. 2003). Likewise, Cr(VI) reduction oc-curred equally rapidly with both Cr(VI)-resistant and plasmid-cured Cr(VI)-sen-sitive strains of P. fluorescens (Bopp and Ehrlich 1988). Chromate resistancedeterminants have been described on plasmids in several bacteria, especially inPseudomonas. However, Cr(VI) reduction determinants have not been found onplasmids. Cr(VI) reduction was independent of chromate resistance, conferredby plasmid pLHB1, in P. fluorescens (Bopp and Ehrlich 1988). In P. mendocina(Bhide et al. 1996), however, plasmid pAR1180 determined both chromate resis-tance and Cr(VI) reduction (Dhakephalkar et al. 1996).

Microorganisms that are known to reduce Cr(VI) reduce it under aerobic oranaerobic conditions, but the physiological role in such transformations is notclear. Earlier reports (Romanenko and Korenkov 1977; Lebedeva and Lyalikova1979; Kvasnikov et al. 1985) have shown that facultative anaerobes (Pseudomo-nas and Aeromonas strains) reduce Cr(VI) to Cr(III) anaerobically. Anaerobicbacteria 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 theelectron acceptor to support the anaerobic growth of bacteria. Enterobacter cloa-cae grew well under aerobic conditions and slowly under anaerobic conditionsat chromate concentrations above 10 mM in nutrient broth, but could reducechromate only under anaerobic conditions (Wang et al. 1989). Also, there isevidence that O2 inhibited the reduction of Cr(VI) by E. cloacae strain HO1 ina medium containing other carbon sources as electron donors (Wang et al. 1989;Komori et al. 1990a, b). Likewise, Escherichia coli could reduce Cr(VI) onlyin the absence of O2 (Shen and Wang 1994a,b).

Under anaerobic conditions, Cr(VI) serves as a terminal electron acceptorthrough electron transport systems involving cytochrome c in Enterobacter cloa-cae (Wang et al. 1989), cytochrome b and d in Escherichia coli (Shen and Wang1993) 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 serveas elecron donors for Cr(VI) reduction. A recent study (Marsh and McInerney2001) established a relationship between the bioavailability of H2 and chromatereduction in anaerobic aquifer sediments. The anaerobic enrichment, developedfrom the sediment, utilized Cr(VI) and was dependent on H2 for growth andchromate reduction. In the absence of Cr(VI), H2 accumulated in the anaerobicmedium. Under Cr(VI)-reducing conditions, however, no H2 and methane accu-mulated because H2 was utilized by the enrichment. When H2 was provided inthe 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 electronacceptor and with a relatively high level of resistance to chromate, have beenisolated from tannery effluents (Basu et al. 1997; Shakoori et al. 1999, 2000).A chromate-resistant gram-positive bacterium (ATCC 700729) tolerated high


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 ashigh as 80 mg/mL, but the reduction required a longer time at 80 mg/mL thanat 20 mg/mL. Chromate reduction occurs either anaerobically (Bopp and Ehrlich1988; 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 bothconditions (McLean and Beveridge 2001) (see Table 1). Agrobacterium radio-bacter EPS-916 (Llovera et al. 1993) and Escherichia coli ATCC 33456 couldreduce Cr(VI) under both aerobic and anaerobic conditions. Likewise, a pseudo-monad, isolated from a wood preservation site contaminated with chromatedcopper 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) faculta-tive anaerobes (five isolates of Aerococcus sp., two isolates of Micrococcus sp.,and one isolate of Aeromonas sp.), isolated from tannery effluent, apparentlyreduced Cr(VI) both anaerobically and aerobically. Cr(VI) reduction by thesefacultative anaerobes in diluted peptone water was more pronounced under an-aerobic conditions (73%–94% reduction) than under aerobic conditions (18%–63% reduction), but conditions used for anaerobic and aerobic systems have notbeen described. Because peptone alone may catalyze chemical reduction ofCr(VI) (Wang and Shen 1995), it was not clear whether the Cr(VI) reduction inmicrobial cultures was caused chemically, microbially, or both. Cell suspensionsof Pseudomonas putida PRS 2000, P. fluorescens LB303, and Escherichia coliAC80 aerobically reduced Cr(VI) to Cr(III) (Ishibashi et al. 1990). Reductionof Cr(VI) in cell suspensions of these bacteria was more rapid and completeaerobically than anaerobically. After disruption of the cells of P. putida andcentrifugation, the supernatant, but not the membrane fraction (pellet), reducedall the added Cr(VI) within 1 hr. Likewise, Wang and Shen (1997) reported thatresting cells of Bacillus sp. and Pseudomonas fluorescens LB300 aerobicallyreduced Cr(VI). However, Cr(VI) reduction by the cells of Escherichia coli wasinhibited in the presence of oxygen (Shen and Wang 1994a,b). Enterobactercloacae, a chromate-resistant strain, could grow in the presence of Cr(VI) underboth aerobic and anaerobic conditions, but Cr(VI) was reduced only anaerobi-cally (Wang et al. 1989). The strain lost both resistance and Cr(VI)-reducingability in anaerobic growth on nitrate.

Cifuentes et al. (1996) reported that organic amendments enhanced the reduc-tion of Cr in soils by indigenous microflora. Generally, Cr(VI) reduction by grow-ing bacterial cells has been demonstrated in media containing natural aliphaticcompounds, amino acids, and fatty acids as electron donors (Wang and Shen1995). Microbial reduction of Cr(VI) occurred during anaerobic degradation ofbenzoate (Shen et al. 1996). A dissimilatory metal-reducing bacterium, Shewa-nella oneidensis, could reduce Cr(VI) when grown on fumarate or nitrate as anelectron acceptor and lactate as an electron donor (Viamajala et al. 2002a).Cr(VI) reduction under fumarate- and denitrifying conditions, dependent on the


physiological state of the culture, was possibly inducible under anaerobic condi-tions. Cr(VI) reduction in the anaerobically grown stationary phase of this bacte-rium is a complex process, possibly involving more than one pathway (Viama-jala 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 tol-uene can also serve as electron donors for Cr(VI) reduction in coculturescontaining E. coli ATCC33456 and P. putida DMP-1 (Shen and Wang 1995).Metabolites produced during phenol degradation by P. putida served as electrondonors for Cr(VI) reduction by E. coli. Technology using such cocultures wouldhelp to simultaneously detoxify both organic pollutants and the toxic Cr(VI).

Nonmetabolizing resting cells of bacteria could reduce Cr(VI), but only inthe presence of an added carbon source (Bopp and Ehrlich 1988; Shen andWang 1994b; Philip et al. 1998). Killed resting cells could not cause Cr(VI)reduction (Shen and Wang 1994b; Wang and Shen 1997). Soluble enzymes incell 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 addedelectron donors.

According to very recent evidence, nonmetabolic Cr(VI) reduction can occuron bacterial surfaces even in the absence of externally added electron donors inthe medium. Thus, Fein et al. (2002) demonstrated that nonmetabolizing cells ofBacillus subtilis, Sporosarcina ureae, and Shewanella putrefaciens could reducesignificant amounts of Cr(VI) in the absence of externally supplied electrondonors. The Cr(VI) reduction by the bacterial strains was dependent on solutionpH, decreasing with increasing pH, and presumably occurred at the cell walland independent of the oxidation of bacterial organic exudates. Such nonmetab-olizing reduction of Cr(VI) by bacteria in nutrient-poor conditions may be im-portant in the biogeochemical distribution of Cr.

Cr(VI) reduction by microorganisms, known to occur under both aerobic andanaerobic conditions (see Table 1), is a redox-sensitive process (Shen and Wang1994b; Chen and Hao 1996). The ability of washed resting cells of Agrobacter-ium 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 underaerobic conditions on glucose, fructose, maltose, lactose, mannitol, or glycerolas the sole carbon and energy source, exhibited similar redox potentials ofaround −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 orsuccinate, to reduce chromate was associated with relatively high redox poten-tials of −138 to −132 mV. Moreover, resting cells pregrown under anaerobicconditions on glucose had lower redox potentials (−240 mV) and a more pro-nounced chromate-reducing activity than did the aerobically grown resting cellson glucose with a redox potential of −200 mV. Likewise, cells pregrown anaero-bically on chromate as the electron acceptor effected more rapid reduction ofchromate than did the anaeorobically grown cells (−198 mV) on nitrate. Evi-dence suggested a negative correlation between chromate reduction by the rest-


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 tooccur under nitrate-reducing conditions but before iron and sulfate reduction(Marsh and McInerney 2001). Evidently, highly reducing conditions, necessaryfor the reduction of iron and sulfate and methanogenesis, may not be requiredfor chromate reduction.

Abiotic reduction of Cr(VI) has also been demostrated in media rich in nutri-ents containing some reductants, especially under predominantly reduced condi-tions. Thus, even in sterile tryptic soy broth, Cr(VI) was reduced abioticallywith time as a function of redox potential (Turick et al. 1996). Thus, more than50% of 25 µg Cr(VI)/mL added to the tryptic sterile broth was reduced abioti-cally 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 chemicalredox 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 non-enzymatically (indirect). There is considerable literature on the involvement ofCr(VI) reductases in direct reduction of Cr(VI) to Cr(III) by bacteria. In growingcultures with added carbon sources as electron donors and in cell suspensions,Cr(VI) reduction can be predominantly aerobic or anaerobic, but generally notboth. 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 andEhrlich 1988; Shen and Wang 1993; Campos-Garcia et al. 1997; McLean andBeveridge 2001). The Cr(VI) reductase enzyme may be present in the membranefraction 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 inP. ambigua (Horitsu et al. 1987), P. putida (Ishibashi et al. 1990), and a Bacillussp. (Campos et al. 1995), with NADH, NADPH or H2 (Desulfovibrio vulgaris)as electron donors and possible involvement of cytochromes b, c, and d. Mem-brane vesicles of E. cloacae, reduced with NADH and then exposed to Cr(VI),oxidized cytochromes c and b and reduced Cr(VI). Evidence suggested thatcytochrome c548 specifically was involved in the reduction of Cr(VI) by mem-brane vesicles (Wang et al. 1991). In the presence of H2 and excess of hydrogen-ase, cytochrome c3, a periplasmic protein, in the soluble cell-free fraction of thecells in D. vulgaris (Lovley and Phillips 1994) reduced Cr(VI) 50 times fasterthan did the Cr(VI) reductase of P. ambigua with NADH and NADPH as elec-tron donor (Horitsu et al. 1987). Soluble fractions of the cell-free extract, largelycytoplasmic, of a pseudomonad from a wood preservation site reduced chro-mate, 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 anaero-bically grown Shewanella putrefaciens MR-1 was formate dependent with high-est activity in the cytoplasmic membrane (Myers et al. 2000). The Cr(VI) reduc-


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 chromatereductase gene, a novel soluble chromate reductase of P. putida has been firstpurified to homogeneity and characterized, using ammonium sulfate precipita-tion, anion-exchange chromatography, chromatofusing, and gel filtration (Parket al. 2000). The reductase activity was NADH- or NADPH dependent. Theoptimum conditions for the chromate reductase were 80 °C and pH 5.0. Kineticproperties of the enzyme showed Km of 374 µM CrO2−

4 and Vmax of 1.72 µmol/min/g protein. Suzuki et al. (as cited in Park et al. 2000) sequenced the geneencoding the chromate reductase (Suzuki et al. 1992) from P. ambigua. Thegenes encoding the chromate reductase in P. ambigua and P. putida were nothomologous, 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 intermediateduring microbial conversion of Cr(VI) to Cr(III). For instance, the NADPH-dependent Cr(VI) reductase in P. ambigua catalyzed the transitory formationof Cr(V) during conversion of Cr(VI) to Cr(III) (Suzuki et al. 1992). Toxicityof Cr(VI) to microorganisms is probably associated with the transient forma-tion of Cr(V) as an intermediate. Cr(V) is formed also during reduction ofCr(VI) in algal cultures and in reactions with physiological reducing agents suchas NADPH, glutathione, and several pentoses (Shi and Dalal 1990).

In most studies, conclusions on microbial reduction of Cr(VI) were based onits disappearance or accumulation of the Cr(III) [determined as the difference intotal Cr and Cr(VI)] as the reduction product with incubation. The colorimetricdiphenylcarbazide method commonly used in Cr(VI) estimation is not specificbecause 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 theoxidation state of the Cr during bacterial reduction of Cr(VI), however, has notbeen attempted until recently. Daulton et al. (2002) used the electron energyloss spectroscopy (EELS) technique to characterize the oxidation state of Crduring Cr(VI) reduction by Shewanella oneidensis in anaerobic cultures. Trans-mission electron microscopy (TEM) of the cells exposed to Cr(VI) showed thatthe cells were encrusted in Cr-rich precipitates, mostly restricted to the outersurface of the cells. These precipitates, based on analysis by EELS, containedCr(III) or its lower state of oxidation. Myers et al. (2000), using electron para-magnetic resonance (EPR) spectroscopy, confirmed the formation of Cr(V) viaone-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, involvinga biotic–abiotic coupling. For instance, Fe(II) and S2−, produced by microorgan-


isms through dissimilatory reduction pathways, can chemically catalyze severalbiogeochemical processes including Cr(VI) reduction (Lovley 1993; Fendorfand Li 1996). Fe(III), an important electron acceptor for microbial oxidation oforganic compounds (aliphatic and aromatic), is one of the most abundant metalsin the soil. A wide range of bacteria couple the oxidation of organic compoundsand H2 to reduction of Fe(III) and SO4 to Fe(II) and H2S, respectively, underoxygen stress conditions (Lovley 1993); this reaction occurs in submerged ricesoils, 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 mech-anism involing iron reduction. Dissimilar Fe(III) reduction by Shewanella algaATCC 51181, a facultative anaerobic bacterium, under iron-reducing conditionsprovided a primary pathway for chemical reduction of Cr(VI), injected into abioreactor, by microbially induced ferrous ion. However, it has been difficultto differentiate the exact contribution between biological (direct) and chemical(indirect: biotic–abiotic) reduction of Cr(VI) in a soil environment. Evidenceusing Desulfovibrio vulgaris as a model chromate reducer suggests that chemi-cal reduction of chromate by Fe(II) was 100 times faster than that by D. vul-garis, a chromate reducer (Wielinga et al. 2001).

In anaerobic environments abundant in Fe(II), nonenzymatic reduction ofCr(VI) by Fe(II) can be as important as enzymatic Cr(VI) reduction (Masschel-eyn et al. 1992). A facultative anaerobe, Pantoea agglomerans SP1, coupledanaerobic growth on acetate and other electron donors to the dissimilatory re-duction of electron acceptors, Fe(III), Mn(IV), and Cr(VI), but not sulfate (Fran-cis et al. 2000). When Cr(VI) was added to this γ-protobacterium culture withelemental sulfur alone, S0-disproportionation to sulfate and hydrogen sulfideoccurred with concomitant growth of the bacterium and reduction of Cr(VI)(Obraztsova et al. 2002). Likewise, P. agglomerans SP1 grew chemolithoauto-trophically by the S0-disproportionation, coupled to reduction of Fe(III) andMn(IV). S0-Disproportionation that may be widespread in certain anaerobic en-vironments may provide an effective mechanism for attenuation of Cr(VI)through its reductive detoxification.

Sulfate-reducing bacteria (obligate anaerobic heterotrophs) couple the oxida-tion 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 an-aerobic conditions prevail (Losi et al. 1994c; Pettine et al. 1994, 1998), as inflooded compacted soils. Likewise, sulfide produced by sulfate-reducing bacte-ria has been implicated in Cr(VI) reduction in marine environments (Smillie etal. 1981). Hydrogen sulfide produced in acid sulfate soil under reducing condi-tions is easily precipitated as FeS in reduced soils (Ponnamperuma 1972) andsediments. Fe(II) (Eary and Rai 1988) and H2S (Pettine et al. 1994), both micro-bially produced, are effective reductants of Cr(VI) under reduced conditions asis the FeS (Patterson et al. 1997). Low concentrations of Cr(VI) can acceleratethe growth and sulfate-reducing activity of sulfate-reducing bacteria (Karnachuk


1995) and thereby the reduction of Cr(VI) by the H2S evolved. A spore-formingsulfate-reducing bacterium, Desulfotomaculum reducens sp. nov. strain MI-1,isolated from sediments with high concentrations of Cr and other heavy metalsby enrichment, could grow with Cr(VI) as sole electron acceptor in the absenceof sulfate with butyrate, lactate, or valerate as the electron donor (Tebo andObraztsova 1998). Cr(VI) was presumably reduced to Cr(III) as Cr(OH)3. In theabsence of Cr(VI), no bacterial growth was noticed.

Biologically generated sulfur compounds with high reducing power such assulfite, thiosulfate, and polythionate can catalyze the chemical reduction of Cr(VI).Chemoautotrophic bacteria belonging to the Thiobacilli group, which can deriveenergy from the oxidation of inorganic sulfur compounds during sulfur oxida-tion, generate a range of sulfur compounds such as sulfite and thiosulfate withhigh reducing power that can in turn catalyze the reduction of Cr(VI). For in-stance, Thiobacillus ferroxidans grown on elemental sulfur has been used toreduce 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 shakeflasks and in fermentation vessels was related to the generation of protons withhigh reducing power from elemental sulfur (QuiIntana et al. 2001). T. ferroxi-dans could reduce Cr(VI) over a wide pH range (2–8), interestingly with morepronounced reduction at lower pH, associated with increased oxidation of ele-mental sulfur to products with high reducing power. Cr(VI) reduction, mediatedby T. ferroxidans in the presence of elemental sulfur, occurred under both aero-bic and anaerobic conditions, but more effectively under aerobic conditions.Evidently bacterial reduction of Cr(VI), involving biotic–abiotic coupling, canoccur 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 ofusing 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), directoxidation of Cr(III) to Cr(VI) in pure cultures of microorganisms has not beendemonstrated yet. Because Cr(VI) is toxic to most microorganisms because ofits high oxidizing nature, direct oxidation of Cr(III) by microorganisms may notbe a predominant process. However, indirect involvement of microorganisms inthe oxidation of Cr(III) to more toxic Cr(VI) through biotic–abiotic coupling isof great concern in environments rich in easily reducible forms of Mn.

In indirect involvement of microorganisms in Cr oxidation, microorganismsfirst oxidize Mn(II) to higher states of oxidation, in particular that, in turn, chemi-cally oxidize the Cr(III) generally to Cr(VI). Oxidation of Cr(III) in the soil in-volving biotic–abiotic coupling could involve essentially Mn-reducing micro-organisms that may indirectly facilitate chemical Cr oxidation (see Fig. 2).

A diverse range of Fe(III) and Mn(IV) oxide forms is potentially formed bymicroorganisms to utilize energy during redox reactions and also to prevent the


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 inthe degradation of organic pollutants (Lovley and Phillips 1988) and reductionof Cr(VI), as well as in anaerobic environments.

Although it is speculated that both Fe and Mn redox systems could be in-volved in Cr oxidation reactions, the main oxidants of Cr(III) in soils are Mnoxides (Bartlett and James 1988). Fe-redox systems are probably involved morein the reduction of Cr(VI) to Cr(III) than in its oxidation. This section focuseson Mn systems, as the role of Mn oxides in chemical oxidation of Cr in soils iswell established [see Avudainayagam et al., this volume (Section II.E)]. Thenature and amount of Mn oxides formed in the soils determine the rate of Croxidation. However, oxidation of Mn(II) to Mn(IV) in the soil is mediated es-sentially by microorganisms. The following section discusses the processes ofmicrobial Mn oxidation in soils, mechanisms of Mn oxidation, and the natureof Mn oxides formed, given that such oxides could play a significant role inCr(III) oxidation, especially in long-term Cr-contaminated soils abundant in Mn.

Microbial Oxidation of Manganese Microorganisms are responsible for muchof the Mn(II) oxidation observed in the environment (Bromfield and Skerman1950), 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 itshigher oxides in natural environments such as soils (Leeper and Swaby 1940),freshwater lakes (Tipping et al. 1985), and estuarine waters and sediments (Eden-born 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 sys-tem. Oxidation of Mn(II) to Mn(IV) is an exergonic reaction, yielding approxi-mately −18.2 kcal mol−1 at 1 M concentrations of the reactants, which can beused as an energy source by microorganisms, but also there are many contradic-tory reports (Ghiorse 1984a,b). That this reaction minimizes the toxic effect ofMn(II) by transforming it to insoluble oxides, which cannot enter the cell, is yetanother advantage (Bromfield 1976).

The heterotrophic microbial populations that can effect the conversion ofsoluble 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-associ-ated materials, which enhance autoxidation later. Indirect mechanisms relate tomicrobially promoted changes in the cell microenvironment that later lead tononbiological oxidation of Mn(II) (Greene and Madgwick 1991).

Microbial Mn Oxidizers A wide range of microorganisms exhibit Mn-oxidiz-ing ability in water and soil matrices: Bacillus SG1 (Rosson and Nealson 1982;Mandernack et al. 1995a), Arthrobacter sp. and Leptothrix sp. (Ghiorse 1984b),


Pseudomonas sp. S36 (Nealson 1983), Oceanospirillum sp. (Ehrlich 1982), Ped-omicrobium sp. (Larsen et al. 1999), the white rot fungus Phanerochaete chryso-sporium (Kirk and Farrell 1987), and an isolate of Streptomyces (Bromfield1979). 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 ismediated either enzymatically or by its binding to other extracellular secretions(Table 2). The majority of the Mn-oxidizing systems are extracellular as incultures of fungi and a Streptomyces, wherein the oxidizing factors are diffusedinto the surrounding environment (Bromfield 1979). The Mn oxides formed infungus are accumulated mainly in the hyphae, resulting in black-colored colo-nies. The take-all fungus (Gaeumannomyces graminis var. tritici) has been re-ported to oxidize Mn in soils. Bacillus sp. and Leptothrix discophora have beenextensively studied for their mode of Mn oxidation. In Bacillus SG1, the sporecoats bind and oxidize Mn(II) in a manner similar to whole spores (deVrindet al. 1986).

Greene and Madgwick (1991) have reported that a Pseudomonas strain inassociation with an alga (Chlamydomonas) oxidized 5 g Mn(II) L−1 to disorderedsemipure γ-MnO2 and manganite (γ-MnOOH) as an intermediate. Indirect Mnoxidation, where the changes in pH or Eh of an environment are modified as aresult of metabolism and growth of microorganisms, is well documented. Bacte-ria and fungi have long been recognised to catalyze this type of nonspecificmanganese oxidation (Ehrlich 1976).

Microbial Mn Oxides Although microbial Mn(II) oxidation has been investi-gated extensively, less attention has been given to the characterization of the

Table 2. Microorganisms capable of oxidising manganese.

Optimumtemperature

Organism Source Oxidizing part for oxidation pH Reference

Bacillus sp. Shore sediments Mature dormant 4 °–45 °C 7.8 Rosson andSG1 spors Nealson 1982

Leptothrix dis- Freshwater sedi- Protein in 7.8 °–28 °C Ghiorse 1984bcophora ments sheaths (110

kDa)Pseudomonas Marine sedi- Extracellular Nealson 1978

sp. S-36 ment enrich- glycocalyxment

Pedomicrob- Freshwater Extracellular Larsen et al.ium sp. enzymes 1999

Scenedesmus Freshwater Extracellular 25 °C 7.9 Knauer et al.sp. 1999


microbially formed Mn oxides (Greene and Madgwick 1991). Generally, theidentification of Mn oxides is problematic because of their complex nature. Themost common methods used for characterization of Mn oxides are X-ray diffrac-tion 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 Lind1983; deVrind et al. 1986) in a two-step process:

3 Mn2+ + 3H20 + 1/2 O2 > Mn3O4 + 6H+

Mn3O4 + 4 H+ > MnO2 + 2 Mn2+ + 2 H2O

Scanning electron microscopy (SEM) studies have showed that microbial Mnoxides are not highly crystalline and are amorphous, recrumpled, and sheetymicrocrystalline solids. Prolonged incubation of cultures for a few weeks ormonths resulted in more crystalline forms of Mn oxides (Mandernack et al.1995b). However, Murray et al. (1985) reported that even after 8 mon the reac-tion did not proceed beyond γ-MnOOH. The reaction was also rapid in initialstages and, once the cells and sheaths were covered with an excess of Mn ox-ides, autoxidation predominated over bacterial oxidation. The morphology ofMn oxides formed in pure cultures of bacteria was very similar to that of manynaturally occurring Mn precipitates surrounding microbes in environmental sam-ples (Ghiorse 1984b).

Factors Governing Mn Oxidation and Oxides Formed The type of oxideformed can vary according to the type of microorganisms and changes in chemi-cal, physical, and growth conditions of cultures (see Table 3). Metallogeniumcultures catalyzed the deposition of disordered Mn oxides such as vernadite(δ-MnO2) in association with a Mn-oxidizing fungus (Emerson et al. 1982). Ingeneral, low Mn oxide concentration (nM–µM) and low temperature promoteddirect 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 Hausmannite, manganite Mandernack et al. 1995bLeptothrix sp. Hausmannite, manganite Adams and Ghiorse 1988Pseudomonas sp. Manganite Nealson et al. 1988Pseudomonas in association Manganite γ-MnO2 Greene and Madgwick 1991

with Chlamydomonas sp.Pedomicrobium sp. Manganite Larsen et al. 1999


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

Tolerable Cr(VI)Organism concentration (mg L−1) References

Pseudomonas K21 5356 Shimada 1979Pseudomonas fluorescens >400 Bopp 1980Arthrobacter sp. 450 Coleman and Paran 1983Agrobacterium sp. 100 Coleman and Paran 1983Escherichia coli 66 Thompson and Watling

1984Frankia strains 52–91 Richards et al. 2002

Mn Oxides in Soil In soils, the commonly occurring Mn oxides are birnessiteand vernadite (McKenzie 1989). Birnessite is mainly formed in neutral and alka-line soils whereas in acid soils the coprecipitates may be manganites. In floodedsoils, the intermediate oxides Mn(Fe)2O3 (bixbyite), 3(MnFe)2O3 �MnSiO3 (braun-ite), (Mn+2Fe)(Mn+3Fe)2O4 (jacobsite), and Mn3O4 � Fe3O4 (vrendenburgite), andperhaps γ-MnOOH (manganite) and Mn3O4 (hausmannite), may be present.When the flooded soils are drained, coprecipitates of iron and manganese areprobably formed (Ponnamperuma et al. 1969). Most of the Mn oxides presentin soil, especially manganate and birnessite (the highly reactive forms of Mnoxides in Cr oxidation), are also produced by microbial cultures (Section II.E).

Biogenic Mn Oxides Responsible for Cr(VI) in Long-Term Tannery Waste-Contaminated Soil In long-term tannery waste-contaminated soil at the MountBarker site in South Australia, Cr(VI) was detected in surface and subsurfacesoil samples and in groundwater water samples (collected in piezometers below50 cm) at levels far above the permissible level of 0.05 mg (kg−1 or L−1), even25 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 >3.5mg Cr(VI) kg−1 soil (Kamaludeen 2002). It appeared that there was no apprecia-ble attenuation of Cr(VI) at this site with time. This finding was surprisingbecause the highly contaminated soil at the waste disposal site was rich in or-ganic carbon (15.7%), derived from animal hides containing electron donorsthat normally should enhance the reduction and thereby also the attenuation ofCr(VI). The contaminated soil contained 0.3–0.6 mg Mn kg−1, mainly as insolu-ble Mn oxides, and a high population of Mn oxidizers (4.7–5.4 × 103 colony-forming units, CFU) in the surface soil (Kamaludeen 2002).

A close correlation existed between concentration of total Mn in the soil andthe concentration of Cr(VI) in soil solution. Although Mn oxidation in soil isknown to be essentially mediated by microrganisms, there was no convincingevidence to suggest that biologically produced Mn oxides catalyzed the forma-tion of Cr(VI) in the contaminated soil. In a most recent study, Kamaludeen


(2002) found that Mn-enrichment cultures, prepared from long-term tannerywaste-contaminated soil at the Mount Barker site, and a bacterium isolated fromthis enrichment culture, could oxidize the Mn(II) to Mn(IV) that, in turn, oxi-dized Cr(III) to Cr(VI). This finding provided probably the first evidence forthe 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 theoxidation of Cr(VI) would explain why natural attenuation of Cr(VI) is nottaking place at the Mount Barker site even in the presence of electron donorsthat should enhance natural attenuation.

IV. Implications of Chromium Transformationson Microorganisms and their Activities

Generally, Cr exists in the environment in stable oxidation states, Cr(III) andCr(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 toxicthan relatively less soluble Cr(III). In alkaline soils, Cr(VI) in solution is thedominant species. In soils at around pH 5.5, however, Cr(VI) complexes withorganic matter and is reduced to Cr(III). Cr(III) is then readily precipitated inacidic soil as insoluble oxides and hydroxides and is hence less bioavailable andless ecotoxic than Cr(VI).

The impact of heavy metals such as cadmium, lead, copper, and zinc onmicroorganisms and their activities in soils has been more extensively studied(Doelmann and Haanstra 1986; Vig et al. 2002) than that of Cr. Previous re-search on the toxicological effects of heavy metals has focused mostly withsoils contaminated over a long term with sewage sludge generally containingmultimetals and complex organic contaminants and to some extent with soilsfreshly spiked with individual or mixed metals. Short-term exposure to contami-nants as in freshly spiked soils causes acute, probably temporary, toxicity. Inlong-term contaminated soils, chronic effects of the pollutants with a persistentshift in microbial populations can be common as a result of elimination of sensi-tive microorganisms coupled with selective stimulation of microorganisms thatare already tolerant or resistant to the pollutant or have evolved by adaptationover a long duration of exposure.

A. Microorganisms

Chromium(VI), which can easily diffuse across the cell membrane in prokary-otic and eukaryotic organisms, is reduced to Cr(III) inside the cells with somereports on transitory formation of Cr(V) and Cr(IV) as toxic intermediates (Ar-slan et al. 1987; Liu et al. 1995). On the other hand, entry of Cr(III) into thecells is restricted due to its precipitation as hydroxides. Hence, Cr(VI) is moretoxic to microorganisms than Cr(III) (DeFlora et al. 1984). Many microbial cellsshowed negative response on contact with Cr. Genotoxic effects in microbialcells are mainly impacted by Cr(VI), resulting in frameshift mutations (Petrilli


and deFlora 1977) and lethal DNA damage (Ogawa et al. 1989). Accumulationof Cr resulting in cell sequestering to combat its inhibitory effect is a majorphenomenon in many microbes resistant to Cr. Prokaryotes are more resistantto Cr than are eukaryotes. Thus, microbes exhibit different Cr tolerance levels(see Table 4). Some of the important changes in microorganisms caused by Crare summarized in Table 5.

Fungi Cr–fungi interactions, mostly related to chromate resistance in filamen-tous fungi and yeasts and chromate reduction by yeasts, have been extensivelystudied (Cervantes et al. 2001). Generally, fungi are less sensitive than bacteriato metals (Doelman 1985). Cr compounds at environmentally relevant concen-trations (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 isolatedfrom Cr-contaminated [long-term or short-term (spiked with Cr)] soil environ-ments. Laboratory induction through enrichment in Cr-amended soil suspensionsor repeated transfers of fungal strains in Cr-supplemented media and mutagene-sis by UV irradiation or chemical mutagens have also been used for isolation ofchromate-resistant fungi. Yeasts resistant to chromate include Candida albicans,Schizosaccharomyces pombe, and Saccharomyces cerevisiae. Chromate resis-tance in fungi may have several reasons: decreased uptake of Cr(VI) (Czako-Ver et al. 1999), defect in sulfate transport (Lachance and Pang 1997), involve-ment of vacuolar structures (Gharieb and Gadd 1998), reduction of Cr(VI), andpresence of acid phosphatase (Raman et al. 2002). Decreased uptake of Cr(VI)is the major mechanism of chromate resistance in both filamentous fungi andyeasts. Biological reduction of Cr(VI), as in many bacteria, can also be impor-tant in the chromate resistance of yeasts, but not in filamentous fungi.

Mycorrhizal Fungi There is considerable interest in exploiting the potential ofmycorrhizal fungi in afforestation and reclamation of degraded lands, minespoils, and metal-polluted soils. For instance, trees inoculated with ectomycorr-hizal fungi become established better than trees without mycorrhizal associationin metal-polluted soils (Brown and Wilkins 1985; Jones and Hutchinson 1986).Metal-tolerant mycorrhizal fungal strains, developed in the laboratory by re-peated transfers in a metal-containing medium and then used for inoculation,have an advantage over sensitive strains in forming an effective association withhost trees in metal-polluted soils. A survey on the distribution of vesicular-arbuscular mycorrhizal (VAM) fungi in tannery waste-polluted soils at threesites in Tamil Nadu, India (Raman and Sambandan 1998) revealed the occur-rence of 15 species of VAM fungi in the polluted soils. Of the 22 plant speciesfrom the polluted sites screened, 19 plant species harbored a mycorrhizal associ-ation. Glomus fasciculatum, G. geosporum, and Gigaspora gigantea were thedominant VAM fungi in the tannery waste-polluted soils with a high concentra-tion of total Cr [1400–1800 mg/kg soil; Cr(VI) level not provided]. The trees


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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 impli-cated in imparting heavy metal resistance to mycorrhizal fungi. The resistancemechanism involved the precipitation of the metals by HPO4 released by acidphosphatase, followed by the binding of the precipitated metal to the cell sur-face. Of the ectomycorrhizal fungi, Laccaria laccata and Suillus bovinus, theformer, which produced more acid phosphatase, was more tolerant to high con-centrations of Cr(VI) (Raman et al. 2002). The compatibility of mycorrhizalfungi with the host plant together with their tolerance to the metal would dictatethe successful establishment of the plant in metal-polluted environments.

Algae The interactions between Cr and algae, terrestrial or aquatic, have beenless intensively studied than Cr–bacteria and Cr–fungi interactions as is the casewith other metal and organic pollutants. There are reports of tolerance or resis-tance of a limited number of algae to Cr, depending on its speciation, but themechanism(s) involved in algal resistance are not understood. Reduction ofCr(VI) to Cr(III) and decreased uptake of Cr by algal cells are not probablyinvolved in algal resistance to Cr. Sequestration of Cr by its complexation withorganic compounds in algal exudates is a possibility, but needs confirmation.Interference in sexual reproduction has been implicated in the evolution of Cr-tolerant algae (Corradi et al. 1995). Recently, Megharaj (unpublished data)found total suppression of algal growth in a long-term tannery waste contami-nated 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 ormore. Viti and Giovannetti (2001) examined the impact of Cr concentration onphotosynthetic microorganisms in three soils whose Cr concentration rangedfrom 67 to 5490 mg/kg. Chronically high concentrations of Cr adversely af-fected aerobic photosynthetic microorganisms and drastically reduced the popu-lation (by most probable number technique) of nitrogen-fixing cyanobacteria.Soils polluted with Cr harbored a low population of the cyanobacteria of thegenus Nostoc, and rarely with heterocysts. In soil enrichment cultures with lowCr levels, however, Nostoc dominated and possessed numerous heterocysts.Toxicity of Cr to algae may involve not only Cr(VI) but also Cr(V), becausetransitory accumulation of toxic Cr(V) has been reported in algal cultures ofSpirogyra and Mougeotia (Liu et al. 1995). In a study on the impact of Cr onalgae, total algal counts should be complemented by changes in algal biodiver-sity. Cr-tolerant algal populations increased in river waters receiving toxic levelsof Cr from a paper mill (Sudhakar et al. 1991). It is not always possible todetermine the impact of a metal on the changes in algal biodiversity in soil orwater environments because of the practical difficulty in selecting an appropriatecontrol under field situations, for instance, adjacent to long-term contaminatedpolluted sites.

Garnham and Green (1995) studied the accumulation of chromate ions by aunicellular non-nitrogen-fixing cyanobacterium, Synechococcus sp. PCC 6301,


and a filamentous nitrogen-fixing cyanobacterium with heterocysts, Anabaenavariabilis, and their ability to reduce Cr(VI). Both cyanobacteria accumulatedchromate in the cell walls rapidly, but at a low level, depending on its concentra-tion; biosorption was an energy-independent process. Cyanobacteria are knownto produce and release complex organic ligands that can bind metals (Megharajet al. 2002). During 18-d growth, A. variabilis reduced almost all the addedchromate to Cr(III) in stoichiometric amounts, with 50% of the latter in the cellsand remaining 50% in the medium (Garnham and Green 1995). Synechococcussp. PCC 6301 was unable to reduce Cr(VI). Cr(VI) reduction by A. variabilispresumably occurred in the heterocysts. It may be worthwhile to mention thatchromate reduction by A. variabilis proceeded at a slow rate when compared tothat reported in bacterial cultures.

Bacteria Gram-positive bacteria are more resistant to Cr than gram-negativebacteria (Ross et al. 1981). Chromium(VI), at 10–12 mg/mL, inhibited mostsoil 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 withbacterial strains used, and EC50 values for Cr(VI) ranged from 0.003 to 7000mg/L. Likewise, Cr was toxic to mixed bacterial populations of sewage originin a chemostat (Lester et al. 1979). Hexavalent Cr is toxic and mutagenic tomost organisms including algae and bacteria (Wong and Trevors 1988). In amore recent study (Francisco et al. 2002) using sodium dodecyl sulfate-poly-acrylamide gel electrophoresis (SDS-PAGE) protein patterns and fatty acidmethyl ester analysis, the major group of bacteria, isolated from a Cr-contami-nated activated sludge with total Cr level of 1.3%, belonged to γ-Proteobacteria,exclusively with strains from the genus Acinetobacter. Evidence suggested thatthe 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 totwo soils, and 3 d later the soils were amended with 2% sewage sludge. Theinhibitory effect of Cr(III) on bacterial population was more pronounced in Gleysoil than in light-colored Andosol soil. High toxicity of Cr(III) in Gley soil wasassociated with increased bioavailability (water soluble and CaCl2 extractable).Concomitant with inhibition of bacteria in Cr-treated soil samples, the fungalpopulation increased severalfold over that in control soils not treated with Cr.With regard to metal toxicity, in terms of ED50 (concentration at which thenumber of bacterial colonies from soil dilutions plated on a medium decreasedto 50% of that in control), Cr (ED50 > 100 µmol) was distinctly less toxic thanCd (ED50 < 20 µmol).

B. Effects on Soil Microbial Community

Phospholipid fatty acids (PLFA) are a good indicator of environmental distur-bance. The principle is based upon the fact that different subsets of a microbialcommunity differ in their fatty acid composition. Membrane lipids and their


associated fatty acids are particularly useful biomarkers as they are essentialcomponents of every living cell and have great structural diversity, coupled withhigh 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 organismdies. The PLFA pattern can therefore be viewed as an integral measurement ofall living organisms present in that sample, reflecting both species compositionand relative species abundance (Baath 1989).

Phospholipid fatty acids have been useful in distinguishing the abundanceand structure of microbial communities in soils (Zelles 1999). There are severalreports on the shift in microbial populations in soils contaminated over the shortand long term with Cu, Pb, Zn, and Ni as compared to uncontaminated soils(Frostegard 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 uncontami-nated soils. PCA of PLFA profiles indicated distinct decreases in fatty acidsspecific 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 yrof amendment with dewatered sewage sludge containing multimetals, includingCr [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 severalfatty acids in sludge-amended soils suggested inhibition of several specific pop-ulations of soil microorganisms. However, in sludge-amended soils, counts ofculturable bacteria significantly increased, in contrast to >20-fold decrease indehydrogenase activity (DHA). There has been no report hitherto, however, onPLFA patterns and shift in microbial populations in soils freshly spiked with Cralone or in long-term tannery waste-contaminated soils with high levels of Cras the major pollutant.

Past studies have shown that chronic metal stress affects the microbial com-munity and decreases microbial biomass, activity, and diversity. Most studieson microbial community structure under chronic metal stress, however, havebeen confined to soils treated with sewage sludge-containing multimetals, oftenwith Cr as a minor constituent. Francisco et al. (2002) attempted to establish arelationship between a culturable microbial community [characterized by fattyacid methyl ester (FAME) analysis and SDS-PAGE] and the Cr(VI) resistanceand Cr(VI) reduction ability of the representative strains of each population inactivated sludge (total Cr, 0.197–2.5 ng/L), under chronic Cr stress, from urbanand industrial tannery areas in Portugal. The ability for aerobic reduction ofCr(VI), when examined with 28 strains representative of each FAME clusterand noncluster in nutrient broth containing 1 mmol/L, was not restricted to onespecies or one genus and was widespread in several Proteobacteria subclassesand in gram-positive G + C bacteria. Cr(VI) resistance and Cr(VI) reduction arenot exclusive to a single group, possibly a result of horizontal genetic transferunder selective pressure from chronic Cr contamination. In bioremediation strat-


egies using microbially mediated Cr(VI) reduction, there is a need for a betterunderstanding of the microbial community and the population response underchronic 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 ofCr are mainly governed by speciation and bioavailability rather than by the totalCr concentration. In alkaline soils, Cr(VI) in solution is dominant, resulting inincreased inhibition, but in acidic soils most of the Cr(VI) complexes with or-ganic 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 majorcontaminants, on microbial biomass has been studied at grassland sites receivingmultimetal-rich military waste disposals (Kuperman and Carreiro 1997) ortreated with timber preservatives containing Cu, Cr, and As (Bardgett et al.1994). Combined concentrations of heavy metals distinctly suppressed the mi-crobial biomass (fungal and bacterial) in grassland sites receiving militarywastes (Cr, 42–143 mg/kg) or wood preservatives. In soils polluted with mili-tary wastes, total and fluorescein diacetate-(FDA-)active fungal biomass wasmore sensitive than FDA-active bacterial biomass (Kuperman and Carreiro1997). It was not clear whether the inhibitory effect on microbial biomass insoils polluted with military wastes was caused by heavy metals or increased soilpH. Likewise, Bardgett et al. (1994) reported a greater sensitivity of fungalbiomass, relative to bacterial biomass, in pasture soils polluted with timber pre-servatives. In these studies with multimetals, however, it is difficult to distin-guish the individual effect of Cr.

Dehydrogenase Activity (DHA) Dehydrogenases are essential enzymes, in-volved in oxidoreduction processes, in all microorganisms. Dehydrogenase ac-tivity (DHA) is one of the important parameters widely used to study the eco-toxic effects of metals and organic contaminants. The main advantage of thissimple, but sensitive toxicity assay is that it reflects the overall microbial activ-ity of the active microbial populations in the soil to provide the current statusof soil health. Reports on the effect of Cr on microbial processes of importanceto soil fertility are summarized in Table 6.

Generally, DHA decreases in sewage sludge-amended soils. Because sewagesludge contains a mixture of heavy metals and organic contaminants, it is diffi-cult 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 soilsamended with sewage sludge containing 220 µg Cr g−1 of soil. This increasewas 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


able

6.E

ffec

tof

chro

miu

mon

mic

robi

alpr

oces

ses

inso

ilsan

dpu

recu

lture

sof

mic

roor

gani

sms.

Eff

ects

,pa

ram

eter

mea

sure

d,So

ilty

peC

rco

ncen

trat

ion

(µg/

g)%

inhi

bitio

nR

efer

ence

s

Silt

loam

8.6

—,

CO

2,10

%C

hang

and

Bro

adbe

nt19

811.

31%

C+

1%dr

ysl

udge

and

1%al

falf

aSa

ndy

400

—,

CO

2,17

%D

oelm

ann

1985

pH7.

0,1.

6%O

M,

and

CE

C1–

2pH

6.2,

64%

OM

260

—,

CO

2,15

%L

ight

hart

etal

.19

83pH

6.7,

3.1%

OM

26—

,C

O2,

10%

Lig

htha

rtet

al.

1983

Fore

stsa

ndy

loam

50—

,C

O2,

20%

Skuj

ins

etal

.19

86pH

7.0

Fore

sthu

mus

1000

00,

Nm

iner

aliz

atio

nR

uhlin

gan

dT

yler

1973

pH3–

4.2

Silt

loam

400

—,

Nm

iner

aliz

atio

n,40

%C

hang

and

Bro

adbe

nt19

82pH

6.9

+1%

slud

gean

dal

falf

a10

0—

,ni

trif

icat

ion,

40%

—,

Nm

iner

aliz

atio

n,18

%So

illit

ter,

agri

cultu

ral

260

—,

nitr

ific

atio

nL

iang

and

Tab

atab

ai19

77pH

5.8–

7.8;

2.6%

–5.5

%C

Agr

icul

tura

l,pH

5.8–

7.8

269

—,

nitr

ific

atio

n,81

%L

iang

and

Tab

atab

ai19

782.

6%–5

.5%

CFo

rest

sand

ycl

ay20

0—

,ni

trif

icat

ion,

26%

Skuj

ins

etal

.19

86pH

7.0

50—

,N

fixa

tion,

93%

Skuj

ins

etal

.19

86Sa

ndy

loam

0.32

Cd

+5.

6Cu

+7.

2Pb

+0,

nitr

ific

atio

nW

ilson

1977

pH6.

6,0.

84%

C30

.7Z

n+

0.76

Cr

2.24

Cd

+7.

5Cu

+47

Pb+

—,

nitr

ific

atio

nW

ilson

1977

148Z

n+

15.7

Cr

0.56

Cd

+1.

88C

u+

12Pb

++,

lag

nitr

ific

atio

nW

ilson

1977

37Z

n+

3.9C

rFo

rest

hum

us20

0C

r+

Ni+

Mo

—,

phos

phat

ase,

20%

Ruh

ling

and

Tyl

er19

73pH

3.6–

4.1


able

6.(C

onti

nued

).

Eff

ects

,pa

ram

eter

mea

sure

d,So

ilty

peC

rco

ncen

trat

ion

(µg/

g)%

inhi

bitio

nR

efer

ence

s

Sand

y15

00—

,de

hydr

ogen

ase,

40%

Doe

lman

nan

dH

aans

tra

1979

pH5.

7,2.

8%O

MC

lay

7500

0,de

hydr

ogen

ase

pH7.

5,3.

2%O

MPe

at75

000,

dehy

drog

enas

epH

5.7,

46%

OM

Sand

y39

0–18

80—

,ur

ease

,10

%D

oelm

ann

and

Haa

nstr

a19

86pH

7.0,

1.6%

OM

Sand

ype

at36

0—

urea

se,

10%

pH4.

4,12

.8%

OM

Agr

icul

tura

lso

il15

0—

,de

hydr

ogen

ase,

83%

Rog

ers

and

Li

1985

1.3%

OM

Fore

stsa

ndy

loam

200

—,

urea

se,

28%

Skuj

ins

etal

.19

86pH

7.0

Six

soils

(pH

5.1–

7.8;

269

—,

urea

se,

17%

–50%

Tab

atab

ai19

771.

5%–5

.5%

C)

Silt

loam

1.0

—,

CFU

bact

eria

20%

Zib

ilske

and

Wag

ner

1982

pH6.

0,2.

1%O

M1.

0—

,A

TP,

60%

556

0,al

tere

dfu

ngal

com

mun

ityFi

veso

ils(p

H4.

4–7.

7;1.

6%–

55,

150,

400,

1000

,30

00,

8000

,T

oxic

toar

ylsu

lfat

ase

at6

Haa

nstr

aan

dD

oelm

an19

9112

.8%

OM

)as

CrC

l 3w

eeks

;de

crea

sed

at18

mon

Four

soils

(pH

6.2–

7.0;

2.7%

–13

0as

CrC

l 3—

,ar

ylsu

lfat

ase,

41%

(ave

rage

Al-

Kha

faji

and

Tab

atab

ai19

795.

3%C

)fo

rfo

urso

ilsFi

veso

ils(p

H4.

4–7.

7;1.

6%–

55–2

000

asC

rCl 3

—,

soil

resp

irat

ion

(glu

tam

icH

aans

tra

and

Doe

lman

1984

12.8

%O

M)

acid

assu

bstr

ate)

23%

insa

nd;

5.12

%in

clay


able

6.(C

onti

nued

).

Eff

ects

,pa

ram

eter

mea

sure

d,So

ilty

peC

rco

ncen

trat

ion

(µg/

g)%

inhi

bitio

nR

efer

ence

s

Five

soils

(pH

4.4–

7.7;

1.6%

–55

–800

0as

CrC

l 3—

,so

ilre

spir

atio

nat

3000

µg/

Doe

lman

and

Haa

nstr

a19

8412

.8%

OM

)g,

+at

8000

µg/g

Tw

oso

ils(p

H5.

8an

d6.

4;0.

5%52

0C

rCl 3

—,3

5%in

Gle

yso

il,5%

inA

n-H

atto

ri19

92an

d3.

2%C

)do

sol;

degr

eeof

inhi

bitio

nre

-la

ted

tow

ater

-sol

uble

Cr

Tw

oso

ils(p

H5.

9an

d6.

4)C

r(II

I):

100

—,

CO

2R

oss

etal

.19

81C

r(V

I):

10an

d10

0—

,C

O2

—,

bact

eria

,C

r(V

I)>

Cr(

III)

Four

soils

(pH

5.6–

7.6;

2.6%

–26

0(C

rCl 3)

—,

L-a

spar

agin

ase,

12%

–21%

Fran

kenb

erge

ran

dT

abat

abai

1991

4.7%

C)

Thr

eeso

ils(p

H5.

6–7.

6;2.

6%–

260

(CrC

l 3)0,

amid

ase

Fran

kenb

erge

ran

dT

abat

abai

1981

4.7%

C)

Thr

eeso

ils(p

H5.

8–7.

4;2.

6%–

130

asC

rCl 3

—,

acid

phos

phat

ase

27%

Jum

aan

dT

abat

abai

1977

5.5%

C)

—,

alka

line

phos

phat

ase

33%

13as

CrC

l 3—

,ac

idph

osph

atas

e5%

—,

alka

line

phos

phat

ase

14%

Sand

ylo

amso

il(p

H6.

15;

orga

nic

Adj

acen

tto

rem

edia

llytr

eate

d—

,D

HA

(sig

nifi

cant

inhi

bitio

nSi

ncla

iret

al.

1997

mat

ter

33.6

%)

(with

chro

mat

efl

uori

dew

ood

with

incr

easi

ngle

ache

dso

ilpr

eser

vativ

e)tim

ber

pole

sco

ncen

trat

ions

ofpr

eser

vativ

eco

nstit

uent

s(f

luor

ide

and

to-

tal

Cr)

;ry

em

eal

supp

lem

ent

led

toin

crea

sed

leve

lsof

DH

Ain

cont

amin

ated

soils

Loe

ssso

il(p

H7.

02;

1.12

%C

;C

r(II

I)as

nitr

ate

and

inth

eox

y-T

oxic

ityto

DH

A:

base

don

tota

lW

elp

1999

15.2

%cl

ayan

ion

Cr(

VI)

asa

Ksa

lt;do

se(E

D50

):H

g>

CU

>8–

12ge

omet

rica

llyin

crea

s-C

r(V

I)>

Cr(

III)

>C

d>

Pb;

ing

dose

sba

sed

onso

lutio

nco

ncen

tra-

tion

(EC

50):

Hg

>Pb

>C

u>

Cd

>C

r(II

I)>

Cr(

VI)


able

6.(C

onti

nued

).

Eff

ects

,pa

ram

eter

mea

sure

d,So

ilty

peC

rco

ncen

trat

ion

(µg/

g)%

inhi

bitio

nR

efer

ence

s

Chr

omiu

m-c

onta

min

ated

activ

ated

Tot

alC

r,0.

197–

2.5

ng/L

Atte

mpt

edto

esta

blis

ha

rela

-Fr

anci

sco

etal

.20

02sl

udge

(chr

onic

stre

ss)

tions

hip

betw

een

cultu

rabl

em

icro

bial

com

mun

ity(F

AM

Ean

alys

isan

dSD

S-PA

GE

)an

dC

r(V

I)re

sist

ance

and

redu

c-tio

nun

der

chro

nic

Cr

stre

ssFu

ngi

Lac

cari

ala

ccat

a,Su

illu

sbo

vinu

sC

rO3

(7.8

,15

.6,

23.4

,39

.0,

and

Invi

tro

grow

thof

both

fung

iR

aman

etal

.20

02(e

ctom

ycor

rhiz

alfu

ngi)

47.0

stim

ulat

edat

0.15

mM

Cr

but

inhi

bite

dat

0.15

mM

;ac

idph

osph

atas

eac

tivity

ofbo

thfu

ngi

incr

ease

dat

all

Cr

con-

cent

ratio

ns;

alka

line

phos

pha-

tase

incr

ease

din

L.

lacc

ata

and

decr

ease

din

S.bo

vinu

sbe

caus

eof

Cr

Ster

eum

hirs

utum

Gro

wth

inhi

bite

dby

Cr(

VI)

,bu

tB

aldr

ian

and

Gab

riel

1997

only

at52

mg/

LP

olyp

orus

cili

atus

,St

ereu

mC

r(V

I)as

K2C

r 2O

7fr

om5

to20

Gro

wth

and

Mn

pero

xida

seac

-Y

onni

etal

.20

02hi

rsut

umm

g/L

tivity

inhi

bite

dby

Cr(

VI)

at10

mg/

L;

noto

xici

tyto

Mn

pero

xida

seon

expo

sure

toC

r(V

I)af

ter

lag

phas

eof

my-

celia

lgr

owth

OM

:or

gani

cm

atte

r;C

FU:

colo

gy-f

orm

ing

units

;D

HA

:de

hydr

ogen

ase

activ

ity.


the amount of metal available to the microbes in soil solution and thereby itseventual effect on DHA. Likewise, moisture at field capacity might mask theeffect of heavy metals on DHA.

Long-term incubation of soils with heavy metals may have a great impact onDHA. However, in general, Brendecke et al. (1993) found that DHA and soilrespiration were little affected by sewage sludge containing multimetals evenafter 4 yrs of its application. However, Kelly et al. (1999a) reported the inhibi-tion of DHA in a soil 18 yr after application of a sludge containing multimetalsincluding total Cr (512 mg/kg). A decreased toxicity of the metals was observedin most cases as the exposure time increased, which was attributed to the elimi-nation of the sensitive microbial populations by the chronic effects of the heavymetals with a concomitant shift toward the dominance of tolerant microorgan-isms. The increased abundance of tolerant organisms in polluted environmentscan be caused by genetic changes, physiological adaptations, or replacement ofmetal-sensitive species with species that already are tolerant of that heavy metal.Bacterial cultures, e.g., Pseudomonas, could tolerate maximum Cr(VI) concen-trations of about 5356 mg L−1. Thus, a distinct shift in population can occur incontaminated soils, especially under long-term impact. Several techniques suchas phospholipid fatty acid (PLFAs) and denaturing gradient gel electrophoresis(DGGE) have been used recently to determine the microbial populations in ag-ricultural soils and in soils polluted with organics and inorganics. Among thesetechniques, PLFA has been used widely to determine the impact of metals onmicrobial communities in soils.

Wood preservatives can be a serious source of environmental contaminants.For instance, softwood timbers are often treated with a preservative containingCr, Cu, and As as protection against insect and fungal attack. The effects ofsurface runoff from such a treatment plant on biological activities in pasturesoils with low, medium, and high levels of contamination were examined byYeates et al. (1994). Metal content in the soil samples ranged between 47 and739 mg Cr kg−1, 19 and 835 mg Cu kg−1, and 12 and 790 mg As kg−1. Highlycontaminated soil samples in the surface layer contained at least 700 mg eachof Cr, Cu, and As kg−1. Generally, normal microbial processes (DHA, basalrespiration, substrate-induced respiration, nitrification, phosphatase activity)were initially inhibited, especially at higher levels of contamination by the threeheavy metals, DHA was the only activity that was distinctly inhibited evenafter 6 wk. In another instance where creosoted electric poles were treated withchromated fluoride wood preservative to eradicate basidiomycete fungi, negativeeffects on soil DHA were associated with increased soil concentrations ofleached 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 ofpreservative pollutants on DHA.

There are not many studies on the relative toxicity of Cr species, Cr(III), andCr(VI) and their toxicity in relation to other metals to the microbial activities insoil. 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


in the order Hg (2.0) > Cu (35) > Cr (VI) (71) > Cr(III) (75) > Cd (90). Sorptionand solubility data, however, revealed that Cr(VI) was the least sorbed and yetleast toxic to DHA among the metals tested, including Cr(III). Based on solutionconcentrations of metals, the toxicity to DHA, in terms of EC50, followed theorder Hg (0.003) > Cu(0.05) > Cd (0.14) > Cr(III) (0.62) > Cr(VI) (78.1). Onewould expect that least sorbed Cr(VI) with high solubility is more bioavailableand 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 includingCr (total Cr, 31 mg/kg; available Cr, 0.7 mg/kg), had no inhibitory effect onvarious soil enzyme activities (alkaline phosphomonoesterase, phosphodiester-ase, arylsulfatase, dehydrogenase, and L-asparaginase) 3 yr after their applica-tion. In fact, these enzyme activities increased linearly up to 90 t/ha (Giusquianiet al. 1994).

Other Enzymes Among the 21 trace elements (applied at 1300 mg/kg)screened for toxicity to arylsulfatase activity, average inhibition of the enzymeactivity by Cr, applied as CrCl3, in four soils used was 41% over control (Al-Khafaji and Tabatabai 1979). Cr(III) appeared to be the most toxic to enzymeactivity, when assayed within 30 min after metal addition, among the heavymetals screened as follows: Cr > Cd > Zn > Cu ≥ Ni > Pb. Its inhibitory effectdecreased by 10 fold when the metal concentration decreased from 1300 to 114mg/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) toarylsulfatase activity decreased with time between 6 wk and 18 mon. After 18mon, Cr was the least toxic among the metals screened. Likewise, Cr(III) ap-plied at 260 mg/kg inhibited L-asparaginase (Frankenberger and Tabatabai 1991)and phosphatase (Juma and Tabatabai 1977) activities. Soil amidase (Franken-berger and Tabatabai 1981) was less affected by Cr and other heavy metals thanurease, arylsulfatase, and phosphatase. The activities of N-acetylglucosamini-dase, β-glucosidase, endocellulase, and acid and alkaline phosphatases, as werefungal and bacterial biomass, were less pronounced in grassland soils pollutedwith military wastes containing multimetals including Cr than in reference soil(Kuperman and Carreiro 1997). Doelman and Haanstra (1989) developed anecological dose–response model using phosphatase activity to determine theshort- and long-term effects of heavy metals.

Cocontamination of polluted soils and wastewaters with recalcitrant aromaticcompounds and heavy metals can be common and widespread. White rot fungiare known for their ability to mediate the degradation of several recalcitrantaromatic compounds. Lignolytic activity of these fungi is catalyzed by extracel-lular oxidative enzymes such as laccase and manganese peroxidase. Heavymetal-tolerant white rot fungi would have a better scope for remediation ofaromatic compounds in a cocontaminated site. The effect of heavy metals


[Cr(VI), Cd(II), Zn(II), Pb(II), and Ni(II)] on the growth and manganese peroxi-dase activity of two wood-rotting Basidiomycetes, Polyporus ciliatus andStereum hirsutum, has been reported (Yonni et al. 2002). Cr(VI) was inhibitoryto the growth of both these white rot fungi at concentrations of 10 µg/mL andabove, and growth was totally inhibited at 20 µg/mL. According to an earlierreport (Baldrian and Gabriel 1997), mycelial growth of S. hirsutum was likewiseinhibited by Cr(VI), but only at a high concentration of 1 mM. The manganeseperoxidase activity of both Polyporus ciliatus and Stereum hirsutum was inhib-ited 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 (Yonniet al. 2002). In combinations of Cr(VI) with one or two more heavy metals,added at individually subtoxic concentrations, growth of S. hirsutum was totallyinhibited (Yonni et al. 2002). However, the toxic effect of Cr(VI) and otherheavy metals, either individually or in combination, on growth and manganeseperoxidase activity of S. hirsutum was alleviated when the metals were addedafter 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 ofaromatic contaminants in environments polluted with heavy metals as cocontam-inants.

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 effectof Cr(III) on nitrification was noticed in all the three soils used and ranged from59% to 96%. Ross et al. (1981) suggested that Cr(VI) may impact nitrificationin 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 immobi-lization, N mineralization, and nitrification) in silt loam amended with NH+

4-N(100 µg/g), 1% sewage sludge and 1% ground alfalfa (Medicago sativa), Crwas the most inhibitory to the N transformations (Chang and Broadbent 1982).At 400 µg/g, all metals inhibited the three N transformations. Inhibition of Ntransformations by heavy metals during a 2- to12-wk incubation followed theorder Cr(III) > Cd > Cu > Zn > Mn > Pb. No clear relationship existed betweenthe toxicity of heavy metals to N transformations and the metals extracted bywater, KNO3, and diethylenetriaminopentaacetic acid (DTPA).

Nitrogen Fixation Heterotrophic N2 fixation was sensitive to Cr at 50 mg/kgin soil spiked with Cr (Skujins et al. 1986). Likewise, nitrogenase of cyanobacte-ria was inhibited by 50% in soils treated with sewage sludge containing Cr (80mg/kg soil) and five more metals at concentrations well below the Europeanguidelines (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


3-wk incubation although Cr(III) was not toxic to bacterial growth in liquidcultures (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 persistentadverse 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-treatedGley soil, respiration was inhibited by 37% over that in the control. The in-creased toxicity to CO2 evolution in Gley soil was directly related to the in-creased bioavailability of water-soluble Cr. Substrate-induced respiration in soilsis an active index of active microbial biomass. Mineralization of glutamic acidas substrate was used as a parameter for screening the toxicity of six heavymetals including Cr added as chlorides in six soils at concentrations rangingfrom 55 to 2000 mg/g (Haanstra and Doelman 1984). All six metals exerted thestrongest inhibitory effect on glutamic acid mineralization in a sandy soil.

Reports on long-term impact of individually applied Cr on microorganismsand 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 infive Dutch soils (Doelman and Haanstra 1984). In sandy loam, clay, and sandypeat soils, Cr distinctly inhibited soil respiration under both short-term and long-term incubation, irrespective of its concentration, although inhibitory effects par-tially decreased with time. In contrast, Cr stimulated soil respiration at unrealis-tically high concentrations of 8000 µg/g in sand and at 3000 and 8000 µg/g insilty loam soil. It is probable that Cr, as a trivalent cation, facilitated the in-creased availability of organic matter to microorganisms in these soils.

Reductive Dechlorination of Organics Cocontamination of the soil and waterenvironments with inorganic and organic contaminants by both natural sourcesand anthropogenic activities is a common occurrence worldwide. Kuo and Gen-thner (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 of2-chlorophenol, 3-chlorobenzoate, phenol, and benzoate in an anaerobic consor-tium. In general, these heavy metals extended the acclimation periods anddistinctly retarded or totally inhibited the anaerobic dechlorination or biodegra-dation of the selected organic compounds, depending on their concentration.Among the metals used, Cr(VI) and Cd(II) were the most inhibitory to dechlori-nation of 3-chlorobenzoate in anaerobic consortium, with total inhibition at 0.5µg/mL. Likewise, the dechlorination of 2-chlorophenol to phenol was inhibitedby Cr(VI), but to a lesser extent than that of 3-chlorobenzoate, with total inhibi-tion only at 5.0 µg/mL. Phenol, formed from 2-chlorophenol at sublethal con-centration of Cr(VI), accumulated at sublethal concentration (2.5 µg/mL) ofCr(VI), but was biodegraded after acclimation. Biodegradation of added phenoland benzoate was inhibited.

Although water-soluble Cr(VI) is 10–100 times more toxic than Cr(III) to


microorganisms, most of the reports on the impact of Cr on biological activitiesin soils has been concerned with Cr(III). In these studies, not much attentionhas been given to the speciation of Cr after its application as Cr(III). In long-term 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 organicpollutants has been studied extensively in the past two to three decades. Pro-cesses developed for remediation of environments contaminated with chromewastes are more suited for aquatic systems than for terrestrial systems. Tradi-tional methods, used especially for wastewaters, involve chemical or electro-chemical reduction of Cr(VI) to Cr(III), precipitation of the latter, and its re-moval by filtration or sedimentation (Eary and Rai 1988). Chemical methods aregenerally not cost-effective and may themselves generate hazardous byproducts(Fendorf and Li 1996). Microorganisms are capable of altering the redox stateof Cr by reducing Cr(VI) to Cr(III) through direct (enzymatic) or indirect (viairon reduction, sulfate/sulfur reduction, or sulfur oxidation) processes.

A. Remediation Technologies for Wastewater and Solutions

Biosorption Sequestration and immobilization of heavy metals, especially inthe solutions of effluents and wastewater, can be accomplished through biosorp-tion, a passive process of metal uptake, using dead biomass in particular (Gadd2000). Biosorption is essentially a nondirected physicochemical complexationreaction between dissolved metal species and charged cellular components thatinvolves sorption or complexing of metals to living or dead cells. The precipita-tion or crystallization of metals leading to their sequestration can take place ator near the cell. Also, insoluble metal species can be physically entrapped inthe microbially produced extracellular matrix or precipitated in bacterial or algalexudates (Volesky and Holan 1995). Extracellular matrices may consist of neu-tral polysaccharides, uronic acids, hexosamines, and organically bound phos-phates that are capable of complexing metal ions. Metabolically mediated accu-mulation is usually intracellular and linked to the control of plasmid linkedgenes (Shumate and Strandberg 1985).

Yeasts and bacteria as well as algae can effectively sequester metals in solu-tions (Kratochvil and Volesky 1998) because of their metal-binding capabilities.Algae such as Scenedesmus, Selenastrum, and Chlorella are known to bioaccu-mulate metals (Brady et al. 1994). The functional groups present in the cellsand cell walls of fungi and algae can serve as the probable sites for biosorptionof metals. For instance, the amino group of chitin (R2-NH) in the alga Sargas-sum and chitosan (R-NH2) in fungi are probably the effective binding site forCr(VI) (see Kratochvil and Volesky 1998). Functional groups such as chitin and


chitosan, however, seem to contribute only 10% of the metals sequestered bythe biomass.

Biosorption, using especially dead biomass, is a cost-effective technology forremoval of heavy metals, and is as effective as ion exchange, but is yet to beexploited 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 removalof Cr(VI) from wastewater systems in the Metex process (Linde AG, Germany)anaerobic sludge reactor, the Bio-Substrat process (Dr. Furst Systems and BKT-Burggraf, Germany) anaerobic microcarrier reactor, and the Agarkar ResearchInstitute chromate reduction process (Pumpel and Paknikar 2001). Biosorptionis not suitable for detoxification of solid Cr wastes in soils.

Biofilms in Bioreactors Bacterial biofilms have been recommended as an effi-cient means of remediating contaminants in the environment because biofilmsprovide tolerance to desiccation, a high level of pollutants, and other stressfactors. Smith and Gadd (2000) used a mixed culture sulfate-reducing bacterialfilm for reduction of hexavalent Cr. In the presence of lactate as the carbonsource and sulfate, 88% of 500 µmol Cr(VI) was removed from the solutionwith bacterial biofilm as insoluble Cr(III) in 6 hr. Because sulfide, a reductantof Cr(VI), was not detected in the medium and no reduction occurred in uninoc-ulated medium, dissimilatory chemical reduction was not involved in Cr(VI)reduction. Evidently, Cr(VI) reduction in sulfate-reducing bacterial films wasbiologically mediated, presumably by enzymes. It is also possible to recover theinsoluble or precipitated Cr(III) from the bacterial films. There is promise forusing this biofilm technology for detoxification of Cr wastes in a bioreactor.

Immobilized Cells Cells immobilized on polyacrylamide gel can be used foreffective detoxification and removal of metals in solution from effluents in areactor. Intact cells of the sulfate-reducing bacterium Desulfovibrio desulfuri-cans, 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 ofthe gel. Immobilized cells also effected the reduction of other oxidized metals,Mo(VI), Se(VI), and U(VI). Immobilized cells may be useful for detoxificationof Cr(VI) in bioreactors.

Bioreactor Using Living Microorganisms Rajwade and Paknikar (1997) de-veloped an efficient chromate reduction process using a strain of Pseudomonasmendocina MCM B-180 for treatment of chromate-containing wastewater.The bacterial strain used was resistant to 1600 mg Cr(VI)/L and reduced 2 mMchromate [100 mg Cr(VI)/L] in 24 hr. In 20-mL continuously stirred bioreac-


tors containing this bacterium and sugarcane molasses as a nutrient, 25–100mg chromate/L was removed within 8 hr (Bhide et al. 1996; Pumpel and Pakni-kar 2001). Efficiency of this bioremediation process is enhanced by anaerobi-osis.

B. Remediation Technologies for Chromium Wastes in Soils

Traditional and innovative methods to manage Cr(VI)-contaminated soils havebeen reviewed by Higgins et al. (1997). The techniques chosen are mainly basedon the feasibility and cost at that particular location and the concentration ofCr(VI) in the polluted soils. Although the total Cr concentration is important, inremediation technologies utmost consideration is given to Cr(VI) levels becauseof its carcinogenicity and mutagenicity. The guideline for risk-based cleanup ofsoil (USEPA 1996c) is 390 mg Cr kg−1 based on the ingestion pathway and 270mg Cr(VI) kg−1 for human exposure by inhalation (USEPA 1996b). There is nocomparable 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 dependson the (1) size, location, and history of the site; (2) soil characteristics such asstructure, texture, and pH; (3) type and chemical state of the contaminants; (4)the degree of contamination; (5) desired final land use; and (6) technical andfinancial means available.

Advances in understanding the chemistry and toxicity of Cr compounds haveled to efforts to remediate Cr-contaminated soil (James et al. 1997). Some ofthe important techniques used are excavation and disposal, soil washing, soilflushing, solidification (ex situ and in situ), vitrification, chemical and biologicalreduction, and phytoremediation; these have their own advantages and disadvan-tages (see Table 5 in Avudainayagam et al., this volume). Most appropriatetechnology is based on the concentration of Cr(VI) present in the contaminatedsoils, nature of the contamination, feasibility, and cost at that particular location.Of all these methods, bioremediation and phytoremediation have been mostwidely used because they are economical and do not release further wastes orharmful byproducts into the environment.

Remediation Using Fe and Organic Amendments The main aim of currentsoil amendment techniques for Cr removal from soils is to promote irreversiblereduction of Cr(VI) to Cr(III) and its hydroxides. Reduction of Cr(VI) can beachieved by incorporation of important reductants such as divalent iron, organicmatter, and organic acids (James 1996). The Cr(VI) reduction reactions are asfollows:

Reduction with Fe and Fe compounds

Fe + CrO2−4 + 0.5 H2O > Fe(OH)3 + 0.5 Cr2O3

6 Fe2+ + 2 CrO2−4 + 13 H2O > 6 Fe(OH)3 + Cr2O3 + 8 H+


Reduction by organic compounds (e.g., hydroquinone)

1.6 C6H6O2 + CrO2−4 + 2 H+ > 0.5 Cr2O3 + 1.5 C6H4O2 + 2.5 H2O

Irreversible reduction of Cr(VI) by Fe(II) to insoluble Fe-Cr(III) hydroxidesis used as the major remediation strategy for chromate-contaminated soils.Amendments with Fe-bearing minerals along with organics could be effectivelyused for reduction of Cr(VI) and precipitation to Cr(III) complexes. Further detailson remediation of soil Cr wastes using Fe and organic amendments are givenby Avudainayagam et al. (this volume).

Organics, as biosolids or other sludge materials, provide a diverse inoculumof microbes that can enhance Cr(VI) reduction. Conversely, organic sourceshave been used for Cr(VI) reduction extensively as an amendment to aid reduc-tion 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 Cr-contaminated groundwater, with and without alfalfa plants, to effectively reducethe 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 load-ings and contact time with the organic matter. More than 90% of the Cr wasrendered immobile and less than 0.5% was taken up by alfalfa, minimizing thetransport of Cr(VI) to drainage water. Organic amendments are known to en-hance the reduction of Cr in soils by indigenous microflora (Cifuentes et al.1996), directly involving enzymatic action or indirectly via iron and sulfur redoxsystems (biotic–abiotic coupling). In soils rich in dissolved organic carbon, for-mation 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, 2-chlorophenol, and p-cresol are suitable electron donors for Cr(VI) reduction(Shen et al. 1996). Chromium-reducing microbes may then be able to simultane-ously remediate organic contaminants as well.

The success of bioremediation processes mainly depends on the level of Crcontamination, the Cr(VI)-reducing efficiency of microorganisms, the stabilityof Cr(III) complexes formed, and conditions that are not conducive for the for-mation and occurrence of Mn oxides.

C. Bioremediation

Bioremediation has been used as a strategy employing introduced or indigenousmicroorganisms for complete transformation of organic pesticides to harmlessend products such as CO2 and H2O. Likewise, microorganisms can transforminorganic pollutants, not necessarily completely, but to compounds with de-creased 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 exhibitsan exceptional capacity to detoxify Cr(VI) by converting it to less soluble and


much less toxic Cr(III) (see Table 1). This capacity is harnessed in bioremedia-tion technology for Cr(VI) wherein the microbial strains are multiplied to adesired population level and pumped into soil or sediments in reactors to pro-mote Cr reduction. The bioremediation efficiency can be enhanced by supple-ments with organic matter and other nutrients in the water or soil to promotethe growth of the introduced microorganisms. The addition of organic sourcesto the soil can promote the proliferation of indigenous Cr(VI)-reducing micro-organisms as well because Cr(VI) reducers, both aerobic and anaerobic, areubiquitous in the soil environment. Losi et al. (1994b) decontaminated largevolumes of Cr(VI)-contaminated water by passing it through an organic amended(cattle manure) soil. Indigenous soil microorganisms augmented by the organicamendments were largely involved in the reduction of Cr(VI) in the water, fol-lowed by precipitation and immobilization of the Cr(III) formed. In in situ tech-niques, nutrients are pumped along with aeration to promote the Cr reductionby aerobic Cr(VI)-reducing bacteria. Some Cr-reducing bacteria and algae havebeen 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). Bioreactorsare cost-effective and are effective for decontamination of Cr(VI)-contaminatedwastewater. However, success has been limited for large-scale decontaminationof Cr(VI)-polluted complex soils.

For treatment of soils enriched with chromite ore processing residue, a tech-nique involving the use of organic-rich acidic manure along with chrome-reduc-ing 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) leachingfrom the waste is effectively reduced in the organic layer, thereby preventingfurther contamination of groundwater (James 1996; Higgins et al. 1997).

As described by Losi et al. (1994c), bioremediation of Cr(VI)-contaminatedsoil is achieved by either direct or indirect biological reduction. Most of thedirect microbial reduction would be expected on surface soils. In the subsurfacelayers, indirect biological reduction of Cr(VI) involving H2S can be predominantand very effective, especially in situations where in situ stimulation of sulfate-reducing 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 shownsome promise for remediation of Cr(VI)-contaminated soils when applied to ananaerobic bioreactor system (Losi et al. 1994c).

Anaerobic Packed-Bed Bioreactor Anaerobic Cr(VI)-reducing microorgan-isms are known to be ubiquitous in soils (Turick et al. 1996). Anaerobic chro-mate-reducing strains have been successfully used for the reduction and sedi-mentation 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


Fig. 3. Bioremediation of chromite ore processing residue in soil using organics andmicroorganisms (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 com-pounds are suitable electron donors for Cr(VI) reduction (Shen et al. 1996). Con-sequently, chromium-reducing microbes may then be able to simultaneouslyremediate organic contaminants as well.

Outlook for Engineered Microorganisms Cr(VI) reduction by a wide range ofmicroorganisms is of environmental and biotechnological significance. Biore-mediation of chromate-polluted environments often poses two major problems:(1) inability of introduced Cr(VI)-reducing microorganisms to establish andfunction at sites polluted with mixtures of contaminants, and (2) biodegradationrates not adequate to achieve acceptable residue levels within an acceptable timeframe. Several strategies have been proposed to enhance the rates of bioremedia-tion of pollutants in such inhospitable environments. One of the approaches isto develop novel engineered strains with increased Cr(VI)-reducing efficiencyfor such situations. Gonzalez (2002) cloned two bacterial genes encoding differ-ent soluble chromate reductases (class I and class II) that reduce Cr(VI) toCr(III). Each class has several close structural homologues in other bacteria.Five of these proteins, overproduced in pure form, could reduce chromate andquinones. Class II proteins could also reduce nitroaromatic compounds. Effortsare underway to use these genes and proteins directly in bioremediation of chro-mate-polluted environments.


Natural Attenuation Natural attenuation involves in situ physical, chemical,and biological processes to decrease the concentration of a contaminant in theenvironment over time without human intervention (National Research Council2000; Suthersan 2002). Biotransformation plays a major role in the natural atten-uation of several contaminants in long-term contaminated sites. In a long-termtannery waste-contaminated site at the Mount Barker site in South Australia,industrial discharges of the waste ceased about 25 years ago. Analysis of sam-ples revealed almost the same Cr(VI) levels in the soil (around 40 mg kg−1) andwater (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 sitealthough the soil was rich in organic carbon (9.8%–15.7%) and harboredCr(VI)-reducing microorganisms (Megharaj et al. 2003). Incubation of this con-taminated soil without and with added cow manure under saturated conditionsled 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 subse-quently subjected to drying. However, no decrease in the concentration ofCr(VI) occurred in the Mount Barker soil held at 70% water-holding capacityeven in the presence of cow manure. Although Cr(VI) can be reduced by a widerange of aerobic microorganisms (see Table 1), its reduction in the contaminatedsoil occurred under saturated conditions and not at 70% water-holding capacity.Reoxidation of Cr(III) and moisture stress conditions probably explains the lackof natural attenuation of Cr(VI) in the contaminated soil at the Mount Barkersite.

D. Applicability of Phytostabilization to Cr-Contaminated Soil

Given the literature available on phytostabilization of metals, it was evidentthat no attempt has been made to stabilize Cr in soils. Theoretically, exudationof organic compounds by plant roots should stimulate the microbial reductionof Cr(VI) because Cr(VI)-reducing microorganisms are known to use a varietyof organic compounds as electron donors. However, there are not many reportson the use of plants for the reduction of Cr(VI) to Cr(III). In general, severalfactors such as site characteristics and possible risk assessment must be assessedbefore implementing the appropriate technique to the field.

Silene vulgaris, an excluder plant (Bini et al. 2001), effectively reducedCr(VI) to Cr(III) and restricted the less bioavailable fractions of Cr in surfacesoils. In a study on the uptake and translocation of Cr(III) and Cr(VI) in riceplants, Cr(VI) reduction was attributed to the plant–microbe interactions in therhizosphere (Mishra et al. 1997). A rhizosphere with intense microbial activitycan 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. Like-wise, some aquatic plants (MelLytle et al. (1998) and possibly rice (Mishra etal. 1997) have a great potential for in situ remediation of Cr because of theirability to reduce Cr(VI) to Cr(III).


VI. Challenges

As stated by James (1996), the complex chemistry involved in Cr transforma-tions causes unique measurement and regulatory challenges. Remediationbecomes complicated in heterogeneous wastes wherein the transformation reac-tions are rapid and interchanging. Although treatment technologies exist forremediation 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 significantrole in determining the success of the remediation of Cr(VI). One of the majorproblems encountered in using Cr reduction as a remediation option is the Mn-assisted reversible oxidation of Cr(III) to Cr(VI) on a shift in the soil to oxidiz-ing conditions or by natural weathering processes. An indirect role of microor-ganisms in Cr(III) oxidation can be envisaged when microbially produced Mnoxides mediate the chemical oxidation of Cr(III). In this regard, it is necessaryto precipitate and immobilize Cr(III) to forms not available for reoxidation inMn-rich Cr-contaminated soils.

Being both environmentally friendly and cost-effective, bioremediation andphytostabilization techniques are very attractive options for remediation ofheavy metals. Bioremediation using ex situ bioreactors and in situ treatmentapproaches, especially for Cr-contaminated soils (Gadd 2000), has been investi-gated; however, few detailed reports exist on targeting Cr associated with tan-nery wastes, especially in the long-term disposal sites. Future research shouldbe directed toward increasing the stability of Cr(III) formed using long-termcontaminated soils. Phytostabilization techniques, with appropriate vegetationand soil amendments (organic manure, phosphate fertilizer, etc.) for immobiliza-tion of the metals, are yet to be effectively explored for sites contaminated withCr (Ward et al. 1999). Overall, there is a need to include and understand themajor biotic–abiotic mechanisms governing Cr transformation to develop effec-tive remediation technologies for complex Cr-contaminated soils. Some of themajor challenges are addressed in this review, with major emphasis on the effectof tannery waste contamination on soil microbial populations and their activi-ties, the biotic–abiotic interactions involved in Cr oxidation, and the applicabil-ity 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 tolarge-scale contamination of land and water. Generally, Cr exists in two stablestates: Cr(III) and Cr(VI). Cr(III) is not very soluble and is immobilized by precip-itation as hydroxides. Cr(VI) is toxic, soluble, and easily transported to waterresources. Cr(VI) undergoes rapid reduction to Cr(III), in the presence of or-ganic sources or other reducing compounds as electron donors, to become pre-cipitated as hydroxides. Cr(VI)-reducing microorganisms are ubiquitous in soil


and water. A wide range of microorganisms, including bacteria, yeasts, andalgae, with exceptional ability to reduce Cr(VI) to Cr(III) anaerobically and/oraerobically, have been isolated from Cr-contaminated and noncontaminated soilsand water. Bioremediation approaches using the Cr(VI)-reducing ability of in-troduced (in bioreactors) or indigenous (augmented by supplements with organicamendments) microorganisms has been more successful for remediation of Cr-contaminated water than soils. Apart from enzymatic reduction, nonenzymaticreduction of Cr(VI) can also be common and widespread in the environment.For instance, biotic–abiotic coupling reactions involving the microbially formedproducts, H2S (the end product of sulfate reduction), Fe(II) [formed by Fe(III)reduction], and sulfite (formed during oxidation of elemental sulfur), can medi-ate the dissimilatory reduction of Cr(VI). Despite the dominant occurrence ofenzymatic and nonenzymatic reduction of Cr(VI), natural attenuation of Cr(VI)is not taking place at a long-term contaminated site in South Australia, even225 years after the last disposal of tannery waste. Evidence suggests that excessmoisture conditions leading to saturation or flooded conditions promote thecomplete removal of Cr(VI) in soil samples from this contaminated site; butCr(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 environmentswith low natural attenuation capacity resulting from reversible oxidation ofCr(III), bioeremediation of Cr(VI) can be a challenging task.

Acknowledgments

This project was funded by the Remediation of Contaminated EnvironmentsProgram, 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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Manuscript received September 30, accepted October 7, 2002.

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