Developments in bioremediation of soils and sediments polluted with ...

Reviews in Environmental Science and Bio/Technology (2005) 4:115–156 DOI 10.1007/s11157-005-2169-4


Springer 2005

Developments in bioremediation of soils and sediments polluted with metals and radionuclides – 1. Microbial processes and mechanisms affecting bioremediation of metal contamination and influencing metal toxicity and transport Henry H. Tabak1,*, Piet Lens2, Eric D. van Hullebusch2,4,5 & Winnie Dejonghe3 1

US EPA, ORD, National Risk Management Research Laboratory, 26 West Martin Luther King Drive, Cincinnati, OH, USA; 2Sub-Department of Environmental Technology, Wageningen University, Wageningen, The Netherlands; 3Flemish Institute for Technological Research, VITO, Belgium; 4Universite´ de Marne-laValle´e, Laboratoire des Ge´omate´riaux, 77454 Marne La Valle´e, Cedex, 2 – France; 5Universite´ de Limoges, Laboratoire des Sciences de l’Eau et de l’Environnement, 87060 Limoges Cedex, France (*author for correspondence, e-mail: [email protected])

1. Introduction 1.1. Bioremediation Technologies Bioremediation technology uses microorganisms to reduce, eliminate, contain and transform to benign products, contaminants present in soils, sediments, water or air. The last 15 years have seen an increase in the types of contaminants to which bioremediation is being applied, including solvents, explosives, polycyclic aromatic hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs). Now, microbial processes are beginning to be used in the cleanup of radioactive and metallic contaminants. Bioremediation is an alternative to traditional remediation technologies such as landfilling or incineration. Bioremediation depends on the presence of the appropriate microorganisms in the correct amounts and combinations and on the appropriate environmental conditions. Although prokaryotes – Bacteria and Archaea – are usually the agents responsible for most bioremediation strategies, eukaryotes such as fungi and algae also can transform and degrade contaminants. Microorganisms already living in contaminated environments are often well-adapted to survival in the presence of existing contaminants and to the temperature, pH, and oxidation–reduction potential of the site. These indigenous microbes tend to utilize the nutrients and electron acceptors that are available in situ, provided liquid water is present.

The bulk of subsurface microbial populations are associated with both microorganisms and dissolved substances, including contaminants and their breakdown products. Bioremediation works by either transforming or degrading contaminants to non-hazardous or less hazardous chemicals. These processes are called, respectively, biotransformation and biodegradation. Biotransformation is any alteration of the molecular or atomic structure of a compound by microorganisms. Biodegradation is the breaking down of organic substance by microorganisms into smaller organic or inorganic components. Unfortunately, metals and radionuclides cannot be biodegraded. However, microorganisms can interact with these contaminants and transform them from one chemical form to another by changing their oxidation state through the addition of (reduction) or removing of (oxidation) electrons. In some bioremediation strategies, the solubility of the transformed metal or radionuclide increases, thus increasing the mobility of the contaminant and allowing it to more easily be flushed from the environment. In other strategies, the opposite will occur, and the transformed metal or radionuclide may precipitate out of solution, leading to immobilization. Both kinds of transformations present opportunities for bioremediation of metal and radionuclides in the environment – either to immobilize them in place or to accelerate their removal.

116 1.2. Metal contamination The challenge posed by metal contaminants lies in the permanence of their nuclear structures: stable isotopes are virtually indestructible. They may be redistributed, however, as well as interconverted among forms of varying chemical and physical properties; as such metal mobility and mutability are the sources of both problems and of many promising solutions. Most heavy metals exist naturally in the earth’s crust at trace concentrations of just a few parts per million (Bodek et al. 1988), sufficient to provide local biota with trace nutrients, but too low to cause toxicity. Disposal of wastes from metal excavation and processing have concentrated these metals to dangerous levels in some soils and sediments, endangering both wildlife and people and motivating efforts to detoxify the soils. EPA standards rarely require that metal contamination be reduced to background levels, however, with preliminary remediation goals for metals varying from 0.2 to 63,000 ppm based entirely on projected risks (Smucker 1996). Metal contamination of soil is especially problematic because of the strong adsorption of many metals to the surfaces of soil particles. Due to the difficulty of desorbing metal contaminants, some traditional remediation methods simply immobilize metals in contaminated soils, for example, by the addition of cement or chemical fixatives, by capping with asphalt, or by in-situ vitrification. Alternatively, soils are often isolated by excavation and confinement in hazardous waste facilities (Schleck 1990; Trezek 1990). Although rapid in effect, both of these options are expensive ($30–$300 m)3) and destroy the soil’s future productivity (Cunningham et al. 1995). Soil washing and pump-and-treat technologies ambitiously attempt to remove contaminating metals from soils by pumping water or other solvents into soil and extracting the liquid downgradient from the contamination. Metals are then precipitated from the liquid, which is recycled if possible. The success of these methods is severely limited by the slow desorption kinetics of adsorbed metals, with the result that additional agents are often used to promote metal transfer to the aqueous phase. Typical additives are acids, chelates, and reductants, which im-

prove cost effectiveness but may introduce further harmful chemicals (Boyle 1993; EPA 1991). A primary strategy of bioremediation is the use of similar metal-mobilizing agents in conjunction with soil washing, with the advantage that they pose no known environmental threat themselves. Biopolymers have been discovered that bind metals with high affinity and travel relatively unimpeded through porous medium. Certain microorganisms transform strongly adsorbing metal species into more soluble forms, and plants are being recruited that act as self-contained pump-and-treat systems. Other methods employ enzymatic ativities to transform metal species into volatile, less toxic, or insoluble forms. Techniques for soil bioremediation are usually designed to be used in-situ, lowering costs; they avoid the use of toxic chemicals, and in nearly all cases, the soil structure and potential for productivity are preserved. Although the understanding of the processes discussed below is far from complete, intense research in bioremediation is rapidly elucidating the mechanisms involved.

2. Microbial processes affecting bioremediation of metals and radionuclides Bioremediation of metals and radionuclides relies on a complex interplay of biological, chemical and physical processes. A fundamental mechanistic understanding of coupling microbial metabolism, chemical reactions and contaminant transport is beginning to develop as well as how these activities could work together to bioremediate metal and radionuclide pollution Microbes exist in complex biogeochemical matrices in subsurface sediments and soils. Their interactions with metals and radionuclides are influenced by a number of environmental factors, including solution chemistry, sorptive reactive surfaces, and the presence or absence of organic ligands and reductants (Figure 1). Microorganisms can interact with metals and radionuclides via many mechanisms some of which may be used as the basis of potential bioremediation strategies. The major types of interaction are summarized in Figure 2. In addition to the mechanisms outlined, accumulationm of metals by plants (phytoremediation) warrants

117

Figure 1. Abiotic and biotic mechanisms influence the fate of metals (M) in subsurface environments. From left to right: Organic material produced by microorganisms can act as ligands (L) and complex with metals, facilitating transport. These ligands can also be degraded by microbes to relase the metal. Microbes can also directly adsorp/desorb, takeup/excrete, and oxidize/reduce metals. Oxidation and reduction will change the valence state of the metal either up or down (Mn±x). Abiotically, metals can form solution complexes or can complex with the surface of clays (kaolinite–polymer complexes). These mineral complexes can sorb metals either directly to the mineral or to organic or iron oxide coatings. Metals can also be directly oxidized or reduced to difference valence states by the ambient redox conditions or by organic reductants (Lred) or oxidants (Lox). Changing the valence state of metals will affect their sorption, mobility, precipitation, and toxicity (image courtesy of S. Fendorf, Standford University).

attention as an additional established route for the bioremediation of metal contamination. The studies of Salt, Raskin and co-workers (Raskin et al. 1997; Salt et al. 1995; Salt et al. 1998) provide a detailed description of use of plants for (1) phytoextraction; the use of metal-accumulating plants to remove toxic metals from soil; (2) rhizofiltration; the use of plant roots to remove toxic metals from polluted waters; and (3) phytostabilization; the use of plants to eliminate the bioavailability of toxic metals in soils. Recently, comprehensive reviews were published on phytoremediation of metal pollution in soils (Chaney et al. 1997; Saxena et al. 1999; Vangronsveld 2000; van der Lelie et al. 2001; Reeves & Barker 2001; Schwitzguebel et al. 2002; Pulford & Watson 2003). Iron cycling and associated changes in solidphase chemistry have dramatic implications for the mobility and bioavailability of heavy metals and radionuclides. Coupled flow and water chemistry control the rate and solid phase products of iron hydroxide reduction and provide critical information in assessing the reactivity of

reduced environments toward metal and radionuclide contaminants. Bioremediation of soils, sediments, and water contaminated with metals and radionuclides can be achieved through biologically mediated changes in the oxidation state (speciation) of those contaminants – biotransformation. Changes in speciation can alter the solubility of metals and radionuclides, and therefore their transport properties and toxicity. The latter two characteristics can determine bioavailability (see van Hullebusch et al. 2005 for further details). Resistance by subsurface microorganisms to the toxicity of heavy metals is critical for the bioremediation of contaminated subsurface sites. Remedial action depends on actively metabolizing microbes, and these microbes might be inhibited by high concentrations of toxic heavy metals. There are at least three types of microbial processes that can influence toxicity and transport of metals and radionuclides: biotransformation, biosorption and bioaccumulation, and degradation or synthesis of organic ligands that

118

Figure 2. Metal–microbe interactions impacting bioremediation.

affect the solubility of the contaminants. Each offers the potential for bioremediation of metallic and radioactive contaminants in the environment. 2.1. Biotransformation Metal-reducing microorganisms can reduce a wide variety of multivalent metals that pose environmental problems. The heavy metals and radionuclides subject to enzymatic reduction by microbes include but are not limited to uranium (U), technetium (Tc), and chromium (Cr). Direct enzymatic reduction involves use of the oxidized forms of these contaminants as electron acceptors. The oxidized forms of U, Tc, and Cr are highly soluble in aqueous media and are generally very mobile in aerobic ground water, while the reduced species are highly insoluble and often precipitate from solution. Direct enzymatic reduction of soluble U(VI), Tc(VII), and Cr(VI)

to insoluble species has been documented and is illustrated in Figure 3. Extracellular precipitation of metals and radionuclides has been demonstrated in a number of microbial isolates. For example, the precipitation of uranium on the cell surface of the bacterium Shewanella is shown in Figure 4. Metal-reducing organisms reduce uranyl carbonate, which is exceedingly soluble in carbonatebearing ground water, to highly insoluble U(IV), which precipitates from solution as the uranium oxide mineral uraninite. Significant advances have been made in understanding the mechanisms of reduction of Fe(III), U(VI) and Tc(VI) in the subsurface bacterium Geobacter sulfurreducens, using the tools of biochemistry and molecular biology. A surfacebound c-type cytochrome (mass 40 kDa) that is involved in the transfer of electrons to extracellular insoluble Fe(III) oxides has been identified

119

Figure 3. Direct enzymatic reduction of soluble heavy metals and radionuclides by metal-reducing bacteria. Non-hazardous organic compounds, such as lactate or acetate, provide electrons used by these microorganisms. Note, however, that if complexed, the reduced species may become mobile.

Figure 4. Microbes can play an important role in immobilizing radionuclides. This image shows a cross section of the bacterium Shewanella with uraninite precipitated on the cell surface (image courtesy of S. Fendorf, Standford University).

and characterized. This protein is not required for the reduction of U(VI), suggesting that the mechanisms of Fe(III) and U(VI) reduction may be distinct. Another c-type cytochrome (9.6 kDa), found in the periplasm, appears to be required for U(VI) reduction (Figure 5). This protein is able to reduce U(VI) in vitro, and a mutant unable to synthesize the protein was unable to reduce U(VI) efficiently. Surprisingly, Tc(VII) is reduced by yet another mechanism, a periplasmic Ni/Fe-containing hydrogenase that uses hydrogen as the electron donor for metal reduction. Moreover, efficient indirect mechanisms may be important in immobilizing Tc in sediments wherein biologically reduced Fe(III) or U(IV) is able to transfer electrons directly to Tc(VII).

A wide range of bacteria reduces the highly soluble chromate ion to Cr(III), which under appropriate conditions precipitates as Cr(OH)3. A number of Cr(VI)-reducing microorganisms have been isolated from chromate-contaminated waters, oils, and sediments, including Arthrobacter sp., Pseudomonas aeruginosa S128, some anaerobic-reducing bacteria, and even several algae. Laboratory experiments with Hanford Site sediments showed that Cr(VI) concentration in water decreased significantly (>66%) in a month-long incubation in the presence of nitrate and added dilute molasses as an electron donor. Thus, the addition of molasses to vadose zone sediments shows potential to decrease the transport of chromium and nitrate into underlying aquifers. Although some microorganisms can enzymatically reduce heavy metals and radionuclides directly, indirect reduction of soluble contaminants may be possible in sedimentary and subsurface environments, although this has not been demonstrated under natural conditions, to date. This indirect immobilization could be accomplished by metal-reducing or sulfate-reducing bacteria. One approach would be to couple the oxidation of organic compounds or hydrogen to the reduction of iron [Fe(III)], manganese [Mn(IV)], or sulfur [S(IV)] in the form of sulfate, [SO42)]. Iron(III) can be biologically reduced to Fe(II), Mn(IV) to Mn(III), and S(VI) (sulfate) to S(II) (hydrogen sulfide, H2S). The reduced product might then, in turn, chemically reduce metals or radionuclides to yield separate or multicomponent insoluble species (see van Hullebusch et al. 2005 for details). The most reactive of these reduced species are ferrous iron [Fe(II)] and H2S. The latter is

120

Figure 5. Transmission electron micrograph (TEM) showing extracellular and periplasmic U(IV) precipitates formed by enzymatic reduction of U(VI) by the subsurface bacterium Geobacter sulfurreducens. (TEM image obtained by S. Glasauer, Guelph University).

generated by the enzymatic activity of ironreducing and some fermentative bacteria, can reduce multivalent metals such as uranium, chromium, and technetium (Figure 6a). The reduced forms of these metals are insoluble and can either precipitate as reduced oxide or hydroxide minerals, or coprecipitate with Fe(III) minerals that form during the reoxidation of Fe(II). In coprecipitation, elements are incorporated in metal oxide minerals as they precipitate from solution. Sulfate-reducing bacteria also may be stimulated to produce a chemically reactive redox barrier (Figure 6b). Hydrogen sulfide generated by sulfate-reducing bacteria could chemically reduce

the contaminant to a form that would be stable for extended periods of time. A study of biofilms in a zinc and lead mine is a good example of indirect immobilization of heavy metals by sulfatereducing bacteria. Sulfide produced by sulfate reduces in the film scavenged zinc and other toxic metals. X-ray fluorescence microbeam analysis revealed that zinc and small amounts of arsenic and selenium were extracted from ground water and concentrated in biofilms in zinc sulphide precipitates (Figure 7) (Courtesy of K.M. Kenner, Arrgonne National laboratory and K.M. Banfield, University of California, Berkeley). Thus, microbial formation of sulfide deposits drastically decreased the migration of contaminant metals.

Figure 6. Indirect mobilization of heavy metals and radionuclides by (a) metal-reducing and (b) sulfate-reducing bacteria.

121 Manganese (III/IV) oxides, which are common mineral phases in many soils and sediments, are also electron acceptors for metal-reducing bacteria. Manganese oxides are also relatively strong oxidants and can oxidize insoluble, reduced contaminants such as the mineral uraninite (UO2), a common product of microbial uranium reduction. Differences in the solubility of oxidized Mn (insoluble) and U (soluble) challenge bacterium predictions of their biogeochemical behaviour during in situ bioreduction. Results from laboratory experiments with the subsurface bacterium Shewanella putrefaciens CN32 quantitatively reduced U(VI) to U(IV), in the form of a solid UO2. The UO2 was observed in regions external to the cell as well as in the periplasm, the region between the inner and outer membrane of the cell, in the presence of the Mn oxides, the reduced U resided exclusively in the periplasm of the bacterial cells (Figure 8) (Courtesy of J. Fredrickson, of Pacific Northwest National Laboratory). These results indicate that the presence on Mn(III/IV) oxides may impede the in situ biological reduction of U(VI) in subsoils and sediments. However, the accumulation of U(IV) in the periplasm indicates that the

cell may physically protect reduced U from oxidation by Mn oxides, suggesting that extensive reduction of soil Mn oxides may not be required for reductive biomobilization of U(IV). Natural organic matter (NOM) may play a role in the reduction of contaminants such as Cr(VI) and U(VI) in subsurface environments. NOM consists of a mixture of organic compounds with different structures and functional groups. These groups include aromatic and phenolic moieties, carboxylic and heteroaliphatic hydroxyl functional groups, and free radicals. In the presence of a metal-reducing bacterium, NOM effectively mediated the transfer of electrons for the reduction of Fe(III), Cr(VI), and U(VI), although the reduction rate varied among different NOM samples and among contaminants. 2.2. Bioaccumulation and biosorption Microorganisms can physically remove heavy metals and radionuclides from solution through association of these contaminants with biomass. Bioaccumulation is the retention and concentration of a substance within an organism. In bioaccumulation, solutes are transported from the

Figure 7. X-ray fluorescence trace metal microanalysis of sulfate-reducing bacteria. Results of electron (bottom left) and X-ray (top right) microprobe analysis of specific biomineralized zinc sulfide precipitates. The sensitivity of the X-ray microprobe enables identification of arsenic and selenium constituents in the zinc sulfide precipitate on the surface of a sulfate-reducing bacterium.

122

Figure 8. TEM images of unstained thin sections from S. putrefaciens CN32 cells incubated with H2, as the electron donor, and U(VI) in bicarbonate buffer in the presence of Mn oxides such as bixbyite or birnessite exhibited an absence of fine-grained extracellular UO2(s) and accumulation of UO2(s) exclusively in the periplasm (images courtesy of J. Fredrickson of Pacific Northwest National Laboratory).

outside of the microbial cell through the cellular membrane, into the cell cytoplasm, where the metal is sequestered. Biosorption describes the association of soluble substances with the cell surface. Sorption does not require an active metabolism. The amount of metal biosorbed to the exterior of bacterial cells often exceeds the amount predicted using information about the charge density of the cell surface. Scientists have demonstrated that charged functional groups serve as nucleation sites for deposition of various metal-bearing precipitates. Three possible non-reducing mechanisms of actinide-microbe interactions are shown in Figure 9. The example shows sorbed U(VI), which appears to remain in its oxidized state. These include: (1) sorption on cell surface sites; (2) additional surface complexation and precipitation of actinides; and (3) precipitation of actinides with bacterial cell lysates. In Grampositive bacteria, surface complexation occurs between organic phosphate groups in cell surface teichoic acid and U(VI). Uranium(VI)phosphate solids are the least soluble of all the U(VI) solid phases. By contrast, Gram-negative bacteria appear to have a lesser ability to sorb U, possibly because they lack these cell-surface organic phosphate groups. One of the most common surface structures found in both Bacteria and Archaea is a crystalline proteinaceous surface layer called the S-layer. The S-layer appears to attenuate the sorption ability of Gram-positive bacteria.

2.3. Siderophore-mediated uptake by microorganisms In aerobic soils, iron exists primarily as Fe(III), which has low water solubility (10)18) and cannot be acquired as the free ion by soil microbes. To circumvent this problem, microbes produce siderophores, low-molecular-weight chelating agents that bind with iron and transport it into the cell through an energy-dependent process (Figure 10) (John et al. 2001). Experiments have shown that various metals can form complexes with siderophores and that many of these complexes are recognized by cell uptake proteins. Study of siderophore actinides and the uptake of these complexes is an important component in the understanding of how microbes and actinides interact in the environment. Researchers have now demonstrated that a microorganism can take up plutonium by the same mechanism it uses to take up iron. The common soil microorganism Microbacterium flavescens uses siderophores to obtain its nutritionally required iron. Bacteria were incubated with the siderophore desferrioxamine-(DF) bound with either plutonium [Pu(VI)], iron [Fe(III)], or uranium [U(VI), as UO2+2]. Using transport proteins, the cells took up the Pu-siderophore complex (Figure 10), although at a much slower rate than they took up the Fe-siderophore complex; however, they did not take up the U-siderophore complex. Only metabolically active bacteria were capable of taking up

123

Figure 9. Three possible non-reductive mechanisms of bacterial cell surface interaction with U(VI) (Courtesy of H. Nitzche and T. Hazen, Lawrence Berkeley National Laboratory, and S. Clark, Wash. State University).

Figure 10. Siderophore-mediated Pu accumulation by Microbacterium flavescens (John et al. 2001).

the Pu siderophore complexes, just as with Fesiderophore uptake. These discoveries could have wide-ranging implications for future biomediation efforts and for more accurate predictions of how plutonium and other actinides behave in the environment. Siderophore-mediated uptake and transport could be an important pathway for environmental mobility and Pu entry into the food chain.

2.4. Microbes and synthetic organic chelators Organic complexing agents can have a profound effect on the mobility of metals and radionuclides in subsurface environments. Synthetic chelators such as EDTA and NTA can form stable, soluble complexes with heavy metals and radionuclides. These chelators were commonly used as cleaning agents during industrial processing of

124 nuclear fuels and were sometimes co-disposed with metals and radionuclides. Metal–chelate complexes have entered the environment and may migrate in ground water. However, the migration of these complexes can be reduced by the biodegradation of the organic ligand (Figure 11). The resulting free metal ions are likely to absorb to mineral surfaces or to form oxide mineral precipitates that would be less mobile in ground water. The degradation of organic chelators associated with metal or radionuclide contaminants, then, might achieve a desirable immobilization of contaminants in place. Some of these chelators can be degraded by naturally occurring microorganisms. A number of EDTA- and NTA-degrading organisms have been isolated and identified. In one study, microbial degradation of EDTA by the environmental isolate BNC1 was influenced by the complex metal. Cobalt(II)–EDTA, cobalt (III)–EDTA, and nickel(II)–EDTA complexes were not degraded, whereas copper (II)–EDTA and zinc–EDTA complexes were. The genes and enzymes responsible for EDTA and NTA degradation have been identified, and the genes have been cloned and sequenced. All the genes necessary to code for degradation of EDTA and NTA occur together in a ‘‘gene cluster’’. Elucidating the genes and enzymes responsible for EDTA and NTA biodegradation will provide an understanding of the environmental and physiological controls on chelate degradation in bacteria, and provide gene probes for monitoring this process in the environment. Such fundamental research on the mechanisms of enzymatic degradation of synthetic chelators is expected to provide useful information for developing bioremediation strategies. Metal-reducing bacteria also sometimes actually promote the mobilization of insoluble forms of some heavy metals and radionuclices. It has been demonstrated that metal-reducing bacteria can solubilize PuO2, which is insoluble, in the

Figure 11. Immobilization of heavy metals by enzymatic degradation of organic chelators, such as EDTA and NTA.

presence of the synthetic chelator NTA. It is thought that the bacteria reduced the insoluble Pu(IV) to Pu(III), which was then complexed by NTA. This process may provide a means of mobilizing Pu from contaminated soils and sediments, and could be a step in the removal of this highly toxic radionuclide from the environment. However, this approach has not been tested in the field. Organic acids formed by the metabolic activity of microorganisms can lower the pH of the system to values that interfere with the electrostatic forces that hold heavy metals and radionuclides on the surface of iron or manganese oxide minerals. Displacement of cations by hydrogen ions may lead to the solubilization of the surface-associated metal or radionuclide. In some cases, the organic metabolites also serve as complexing agents that can form soluble metal– ligand complexes. These complexing agents (which include dicarboxylic acids, phenolic compounds, ketogluconic acids, and salicylic acids) have been shown to promote the dissolution of a wide range of heavy metals and radionuclides, including PuO2. Therefore, biogenic production of complexing agents can accelerate the movement of metals in soils and sediments.

2.5. Microbial metabolism of iron reducing bacteria 2.5.1. Dissimilatory iron reduction Iron is extremely abundant in the Earth’s crust, primarily in the form of insoluble Fe(III) oxides. The reduction potential of Fe(III)/Fe(II) is electropositive (Table 1). A number of microorganisms are able to couple oxidation of hydrogen or organic compounds to the reduction of Fe(III) and gain energy for growth. The use of iron or other metals as terminal electron acceptors is called dissimilatory metal reduction. (Not all dissimilatory metal reduction, however, is linked to energy conservation.) Geological and microbiological evidence suggests that Fe(III) reduction was a very early form of respiration on Earth. A phylogenetically diverse group of Bacteria and Archaea is known to conserve energy to support growth by oxidizing hydrogen or organic compounds (including contaminants such as aromatic hydrocarbons) with the reduction of

125 Table 1. Microbially significant half-reaction reduction potentials Transformation O2 depletion Denitrification Mn reduction, Mn(IV) to Mn(II) Fe reduction, Fe(III) to Fe(II) Sulfate reduction, S(VI) to S(-II) Methane generation, C(IV) to C(-IV) H2 generation, H(I) to H(0)

Reaction

Eh, Volts (@ pH 7) +

0.5O2+2H fi H2O NO3)+6H++5e) fi 0.5N2+3H2O MnO2+4H++2e) fi Mn2++2H2O Fe(OH)3+3H++e) fi Fe2++3H2O SO42)+10H++8e) fi H2S+4H2O HCO3)+9H++8e) fi CH4+3H2O H++e+ fi 0.5H2

0.82 0.71 0.54 0.01 )0.22 )0.26 )0.41

see Stumm and Morgan (1996).

Fe(III). Such a group includes species from such genera as Geobacter, Desulfuromonas, Pelobacter, Shewanella, Ferrimonas, Geovibrio, Geothrix, and others. These organisms have a broad spectrum of other metabolic capabilities as well. Many dissimilatory metal reducers such as Geobacter species can reduce soluble U(VI) to insoluble U(IV) (Figure 12). Dissimilatory metal-reducing microorganisms might prevent migration of uranium in ground water by precipitation and immobilization in the subsurface. When a simple organic compound such as acetate is added to the subsurface, aerobic microorganisms quickly consume available dissolved oxygen and nitrate. Then dissimilatory

metal-reducing microorganisms begin to metabolize acetate, oxidizing it to CO2 while reducing available metals. While Fe(III) is generally the most abundant metal electron acceptor in the subsurface, dissimilatory metal-reducing microorganisms can also simultaneously reduce U(VI) to U(IV), precipitating it out of ground water. This has been demonstrated conclusively in laboratory studies, and is the basis for new strategies of in situ biomediation. 2.5.2. Mechanics Mechanisms for ion reduction appear to vary among dissimilatory metal-reducing microorganisms (Figure 13). Some species use strategies to

Figure 12. Geobacter sulfurreducens growing with insoluble Mn(IV) oxides as the electron acceptor (Image courtesy of D. Lovley, University Massachusetts).

126 overcome the need for direct contact with Fe(III) oxides. For example, Shewanella oneidensis is a versatile microbe that can use oxygen, nitrate, uranium, manganese, and iron as electron acceptors. This bacterium appears to release quinones into the culture medium during growth that serve as electron shuttles between the bacterium and the Fe(III) oxide. Shewanella alga and Ceothrix fermentans (in addition to producing electron shuttles) solubilize Fe(III) during growth, presumably by releasing one or more Fe(III)-complexing compounds called chelators. By contrast, Geobacter metallireducens does not release electron shuttles and is highly adapted to contact with the solid Fe(III) oxide. When growing on insoluble Fe(III) or Mn(IV) oxides, this microorganism produces flagella and uses chemotaxis to find the electron acceptor. Pili are also produced under these conditions, presumably to attach to insoluble oxides. Although dissimilatory metal reducers are of obvious importance to developing strategies for biomediation of organic contaminants as well as metals and radionuclides, this process can be

slow. One idea for stimulating their activity in aquifer sediments is to add humic acids or other quinone-containing compounds to which Fe(III)reducing microbes can transfer electrons. Electron shuttling via extracellular quinones may accelerate the rate and extent of biomediation. 2.5.3. Exploring the diversity of iron (III)reducing bacteria in subsurface sediments Iron(III)-reducing bacteria are thought to catalyze a large number of sedimentary processes that have important impacts on biomediation. Many new strains of iron-reducing microorganisms have now been isolated from uranium-contaminated subsurface sediments, expanding our knowledge of the diversity of this environment. Gene-sequencing methods have been used to classify these isolates, which include Gram-positive genera (Clostridium, Camoccus) that were not closely related to any previously characterized pure cultures of Fe(III)-reducing bacteria. The Clostridia are well-studied anaerobic bacteria that have been isolated from sediments since the origin of environmental microbiology.

Figure 13. Mechanisms of Fe(III) reduction by dissimilatory metal-reducing bacteria (c = chelator) (Image courtesy of D. Lovley and ASM News).

127 However, all of the Clostridia isolated previously were fermentative organisms incapable of respiration. In contrast, many of the Clostridium strains described from contaminated subsurface sediments were shown to conserve energy for growth by coupling the respiration of Fe(III) oxide minerals to the oxidation of organic acids (acetate or lactate). Several of the bacterial isolates were also shown to reduce U(VI). Although their environmental significance remains to be explored, these newly isolated FeRBs could play an important role in subsurface biomediation.

3. Microbial mechanisms involved in bioremediation of metal and radionuclide contaminated soils and sediments 3.1. Biopolymers and their use in bioremediation of metal contamination Microbially produced macromolecules with metal-binding properties include cyclodextrins, exopolysaccharides and amphipathic molecules termed ‘‘biosurfactants.’’ Through combinations of carboxyl, phosphoryl, and hydroxyl groups, these molecules complex divalent cations, increasing their aqueous solubility to enhance the efficiency of soil washing. Biopolymers may be injected with washing solvents or produced within contaminated soil by metal-tolerant microorganisms. The properties of these molecules are also potentially alterable by genetic engineering to produce structures with higher metal specificities and complexing strengths. 3.1.1. Biosurfactants Composed of polar, often phosphate-based ‘‘heads’’ and non-polar lipid ‘‘tails’’, biosurfactants combine complexation activity with physical sequestration of the complexed ions (Miller, 1995). A ‘‘surfactant,’’ or surface-active agent, diminishes the attractive forces among the molecules of a liquid at the liquid’s surface. Because the free energy of the surface is thereby lowered, surfactant molecules tend to concentrate at such surfaces, including soil–water interfaces (Tinoco Jr. et al. 1985). This characteristic directly benefits the surfactant’s bioremedial purpose by concentrating the molecules near sorbed metals,

enhancing rates of attainment of complexation equilibrium. At a sufficiently high-concentration of 1–200 mg/L, called the critical micelle concentration (CMC), biosurfactant molecules begin to aggregate into micelles (spherical bilayers), sequestering the chelated ions within the particles. These particles usually measure less than 50 nm in diameter and are thus expected to avoid filtration effects in most soils. Adsorption may still impede their progress, however (Miller 1995). The most thoroughly studied bio-surfactant is the anionic rhamnolipid of Pseudomonas aeruginos (Figure 14). The rhamnolipid forms complexes with Cd2+, Zn2+ and Pb2+ in aqueous solution, with stability constants (Log K) of 6.5, 6.6 and 5.4, respectively (Herman et al. 1995). These are far lower than the metal–EDTA constants (b=105.0 to 106.0) suggesting that the biosurfactant should compete successfully with soil organic matter for metal complexation. The rhamnolipid’s metal-mobilizing capability has been investigated in a sandy loam soil of minimal (0.11%) organic matter, showing that lead desorption is the rhamnolipid’s greatest strength. In soil samples with adsorbed cadmium, zinc and lead (total 3.4 mmol kg)1), 12.5 mM rhamnolipd enhanced lead desorption relative to an electrolyte control solution, while 80 mM rhamnolipid was needed to improve cadmium and zinc desorption. The variable effectiveness is likely to be due to properties of the metals: lead is a softer Lewis acid than cadmium or zinc. Therefore, it tends to form more stable complexes and is less affected by ion-exchange processes. These characteristics are evident in both the notorious difficulty of lead removal from soil and in the measured metal– rhamnolipid stability constants. While a fairly concentrated (110–180 mM) electrolyte solution released the harder acids by cation exchange,

Figure 14. Rhamnolipid structure of Pseudomonas aeruginosa ATCC 9027. For C18 monorhamnolipid, m+n=10; for C20, m+n=12; for C22, m+n=12; and for C24, m+n=12. (Reprinted from Zhang & Miller, 1994, with permission. Copyright 1994 American Society for Microbiology).

128 rhamnolipid complexation was required to mobilize soil-sorbed lead. If rhamnolipids are generally effective in mobilizing soft Lewis acids, they will provide a much-needed tool in soil remediation (Herman et al. 1995). After complex formation and presumably after subsurface transport and recovery, the rhamnolipid can be recovered and recycled: acidification of a solution containing a cadmium–rhamnolipid complex precipitates the rhamnolipid, releasing nearly all of the metal into solution. The rhamnolipid may then be separated by centrifugation, redissolved at neutral pH and reused, while the metal is recovered from the supernatant by alkali precipitation or cations exchange (Tan et al. 1994). One limitation to rhamnolipid use is the high concentration needed to release metals from soil, due to strong adsorption of the rhamnolipid itself. Rhamnolipid-soil sorption is, however, greatly diminished in solutions of low ionic strength, suggesting that ‘‘cation bridging’’ between the anionic head group and cations sorbed to the soil occurs. Efforts are currently underway to improve the efficiency of rhamnolipid treatment (Herman et al. 1995). Remarkably, the P. aeruginosa rhamnolipid can also promote the transport of entire microorganisms through a solid matrix, in this case a porous sand. Experimentation and modeling have shown that the rhamnolipid, used at 150–1000 mg l)1, acts by decreasing adsorption of the cells to the particle surfaces, either by increasing the negative charge density of the surfaces and thus the repulsion between cells and surfaces, by dissolving extracellular polymers used in adhesion, or by competing with the microorganisms for sorption sites. Because in-situ bioremediation will sometimes require the delivery of microorganisms to sites of contamination and because most bacteria have low mobility in soils, this rhamnolipid property may become quite valuable (Bai et al. 1997). 3.1.2. Cyclodextrins Cyclodextrins are cyclic oligosaccharides (Figure 15) formed during bacterial degradation of plant starches (Wang & Brusseau 1993). The central ‘‘cavity’’ of relatively low polarity allows formation of 1:1 inclusion complexes with non-polar organic molecules, while the hydrophilic outer

ring enables the complex to remain water-soluble. The addition of carboxylate groups to the outer ring has given the cyclodextrins metal-complexing ability as well: carboxymethyl-b-cyclodextrin, or CMCD, strongly complexes Cd2+ with a constant of 103.66 for 1 g/l CMCD (Wang & Brusseau 1995) and the presence of phenanthrene as a model organic contaminant does not reduce this metalcomplexing capacity (Brusseau et al. 1997). Like biosurfactants, cyclodextrins do lower the surface tension of water, but they do not have a critical micelle concentration (Wang & Brusseau 1993). Cyclodextrins are small enough to avoid pore exclusion in soils and experiments have shown no measurable adsorption of the cyclodextrin to sandy subsoils and surface soils (Brusseau et al. 1994), indicating that cyclodextrin mobility in soils should be good. Further, cyclodextrins are generally non-toxic (Wang & Brusseau 1995). In column experiments, mixtures of carboxylated and non-carboxylated cyclodextrins dramatically improved cadmium, nickel, strontium and phenanthrene elution from soils in comparison to dilute electrolyte solutions, removing approximately 90% of the metals and 86% of the phenanthrene within 30 pore volumes (Figure 16). This ability of cyclodextrins to mobilize metal and organic contaminants simultaneously is a significant attribute, suggesting that use of cyclodextrins may greatly enhance pump-and-treat remediation of sites with mixed metal-organic contaminants. 3.1.3. Exopolysaccharides Exopolysaccharides are relatively large molecules, frequently with molecular weight of 106 daltons or more, with anionic character due to acidic functional groups. Exopolysaccharides are typically highly hydrated and flexible in solution, and as metals are bound, they may fold into more compact structures. Some polysaccharides aggregate into large flocs in the presence of metals, possibly due to formation of insoluble metal hydrolysis products that form nuclei for further precipitation. While useful in the recovery of metals from waters, this reduced solubility greatly diminishes such a polymer’s mobility and bioremedial effectiveness in soils (Brierley 1990). A great diversity of microorganisms produces free and/or capsular exopolysaccharides, and

129

Figure 15. Structure of hydroxypropyl-b-cyclodextrin (Reprinted from Wang & Brusseau (1993), with permission. Copyright 1993 American Chemical Society).

Figure 16. Elution of a southern Arizona surface soil (0.68% organic carbon, 88.5% sand, 4.3% silt, 10.2% clay, pH 7.5) contaminanted with cadmium (1.6 mg/g soil), nickel, strontium and phenanthrene (7.1 lg/g soil) using CMCD (carboxymethyl-b-cylodextrin) or 0.01 M KNO3. Phenanthrene elution is not shown (Reprinted from Brusseau et al. 1997, with permission. Copyright 1997 American Chemcial Society).

130 most of these are known to possess metal-binding activity. The emulsan of Acinetobacter RAG1, for example, binds up to 240 lg UO2 per mg. Other exopolysaccharides produced by Pseudomonas, Arthrobacter and Klebsiella have bound per mg polymer, up to 96 lg uranium, 3.3 lg cadmium and 22 lg copper, respectively (Miller 1995). Recently, the first demonstration has been made that a expolysaccharide can desorb a metal bound to a mineral surface and transport the complexed metal through a porous medium. This structurally undefined polysaccharide, produced by a Gram negative soil isolate Strain 9702-M4, actively complexed Cd2+ and sorbed only weakly to sand particle surfaces. Most importantly, the polysaccharide released 57% of the (approximately 0.08 ppm) cadmium adsorbed to sand particles in 1500 pore volumes, compared to nondetectable cadmium release in polymer-free controls. The polymer-facilitated cadmium transport is an encouraging result, but not necessarily a general one; in particular, it waits testing in more finely textured soils (Chen et al. 1995). 3.2. Microbial leaching Bacteria that leach metals from ores are chemolithotrophic, or ‘‘rock-eating,’’ meaning that they obtain energy from the oxidation of inorganic substances coupled to the respiration (reduction) of oxygen. Chemolithotrophs of the genera

Thiobacillus and Leptospirillum, the most important in microbial leaching, flourish in metal sulfide deposits exposed to air and water. They oxidize iron and sulfides, generating sulfuric acid and releasing associated metals into aqueous solution (Brierley 1982). A model of the indirect leaching mechanism of Thiobacillus ferrooxidans consists of several steps (Figure 17): the bacteria bind themselves to a metal sulfide substrate with extracellular polymers, forming biofilm of one or two cell layers. The exopolymers hold 0.5–5% iron, needed to overcome the repulsion between negatively charged sulfide minerals and anionic cell surface macro-molecules (Blake et al. 1994). The bacteria oxidize the bound iron with molecular oxygen, forming Fe (III) hexahydrates at low pH, and the Fe(III) hexahydrates in turn oxidize mineral-surface sulfides to soluble thiosulfates. Thiosulfates may be further oxidized by T. ferroxidans or by commensal bacteria that cannot oxidize sulfide directly, ultimately producing sulfuric acid (Sand et al. 1995). Thiobacilli are widely exploited: the mining recovery of metals from low grade ores through processes known as ‘‘heap leaching’’ and ‘‘dump leaching ‘‘ in which excavated ores are mounded together, often on an impermeable slope and irrigated with acidified water to promote indigenous, microbial activity. The leachate is collected at the bottom for metal recovery (Brierly 1982). The success of microbiological mining has inspired work in recovering metals from other

Figure 17. A model for the indirect leaching attack mechanism as catalyzed by a metal-sulfide-attached cell of Thiobacillus ferrooxidans, exaggerating the exopolymer layer to show its importance (Adapted from Sand et al. 1995).

131 substrates as well, notably metal contaminated sewage sludge. Such sludges, generated in abundance by municipal water treatment plants, can be spread onto agricultural land for decomposition if their metal contents do not exceed regulatory standards. Unfortunately, approximately 50% of the sludges in the USA and Canada do contain metals, predominantly as metal sulfides, in excess of the standards. Attempts at chemical dissolution of sludge metals, usually by addition of acids have proven costly and ineffective (Couillard & Mercier 1991). To solve the sludge problem, a continuous flow bioreactor system has been developed for sludge treatment, using T. ferrooxidans to catalyze metal dissolution and using industrial FeSO4 Æ7H2O as an energy source for the bacteria. With an hydraulic residence time of 18 h and sludge volume of 30 L, the following metal solubilizations have been achieved; Cu, 91% of 2500 ppm; Zn, 94% of 1800 ppm; Mn, 93% of 380 ppm; Ni, 67% of 23 ppm and Cd, 67% of 16 ppm. After treatment, therefore the sludges easily met the Quebec limits of 1000 ppm Cu, 2500 ppm Zn, 1500 ppm Mn, 180 ppm, Ni and 15 ppm Cd. The biological leaching method reduces the necessary acid input to 15% and the incubation time to 10% of those required in batch-type chemical leaching methods (Couillard & Mercier 1991; Couillard & Mercier 1993). In further experiments, leaching of Pb was greatly improved when the FeSO4 Æ7H2O substrate was replaced with FeCl2 presumably due to the low solubility of lead sulfate. The bacterial population adapted well to the FeCl, but not to the substitution of HCl for H2SO4 in initial acidification of the system, due to either chloride intolerance or to a sulfate requirement. Further research is needed to optimize lead solublization by these methods, but the leaching technique for the other metals appears ready for implementation (Mercier et al. 1996). In the past decade, the ability of Thiobacilli to remove trace metals from soils and sediments has been studied (Tichy et al. 1998). Operational conditions studied include soil sediment characteristics (Loser et al. 2004), initial pH (Chen & Lin 2001a), percentage of inoculum, retention time, initial solid content (Chen & Lin 2000) and nutrition (Chen & Lin 2001b). When insoluble elemental sulfur is used as substrate in the

bioleaching process, the microbial oxidation of the sulfur by Thiobacilli is believed to take place by the adsorption and growth of bacteria on the surface of the sulfur particles (Boseker 1997). The adsorption of bacteria to the solid substrate and plays a vital role in the bioleaching process the sulfur oxidation rate (Chen & Lin 2001b). Suspension leaching is not economically feasible for treating large amounts of sediment. An alternative is solid bed leaching, as done with ores. In a solid bed, the density and particle size distribution of the solid matrix are important factors regulating mass transport, sulfur oxidation and microbial activity. Improving the physicochemical properties and the structure of dredged sediments by pretreatment with plants enabled leaching of heavy metals in laboratory percolator systems with nearly the same efficiency as suspension leaching (Loser et al. 2001). Nevertheless, good performance in the laboratory does not guarantee similar efficiency on a large scale since important process parameters can be carefully controlled in the laboratory and fluctuations are usually avoided. Some influences on process operation under practical conditions, such as the effects of solid inhomogeneity, channeling, or pH and temperature gradients can only be studied reliably on a pilot scale. To be applicable in a commercial plant, the process must be scaled up without loss of efficiency (Seidel et al. 2004). 3.3. Dissimilatory metal reduction Respiratory microorganisms obtain energy from the enzymatically controlled oxidation of electron donors, such as glucose, acetate, or H2, coupled to the reduction of electron donors, such as O2, NO3), Fe(III) and SO42) (Table 2). Recently, several toxic metals and metalloids have joined the list of terminal electron acceptors: As(V) as arsenate, Se(VI) as selenate and U(VI) as uranyl acetate (Lovley et al. 1991; Ahmann et al. 1994; Oremland et al. 1994). In addition chromium(VI) and plutonium(IV) appear to be reduced by respiratory enzymes, although respiratory growth with these acceptors has not been proven (Lovley and Phillips 1994; Rusin et al. 1994). The importance of these reductions for bioremediation lies in the profound physical and chemical changes that the metals exhibit following reduction, creating

132 opportunities for detoxification or for separation of the metallic compounds from soil matrices. 3.3.1. Selenium reduction Selenium contamination has gained particular attention in the Central Valley of California and other areas of the western United States where naturally selenium-rich soils are irrigated for agriculture. Excess water is drained from the fields and carried to evaporation ponds, some in ecologically sensitive areas such as the Kesterson Wildlife Refuge, where the selenium has become concentrated enough to cause toxicity, deformities, reproductive problems and death in waterfowl (Oremland et al. 1989; Macy 1994). The selenium problem has inspired investigations into two bioremedial strategies: microbial reduction of soluble selenate into insoluble elemental selenium, discussed below, and microbial methylation of inorganic species to form volatile products, considered later in this chapter (see 3.4.1). Selenium is a trace metalloid with a redox chemistry similar to that of sulfur, possessing primary oxidation states of VI, IV, O and –II. Selenate (SeO42)) is the form found most commonly in soil porewaters (Frankenberger Jr. & Karlson 1994a). Because the reduction potential of selenate (Se(VI)) to elemental selenium is sufficiently high, researchers hypothesized that selenate reduction coupled to hydrogen or acetate oxidation could generate enough energy to support

anaerobic microbial metabolism (Table 2). Three selenate-respiring bacterial strains have since been isolated from selenate-rich-sediments: SeS, SES-3 and Thauera selenatis (Oremland et al. 1994, 1989; Macy et al. 1993). Bacillus and Microbacterium species have also been identified that reduce, but do not necessarily respire, selenate (Combs et al. 1996). Selenate and nitrate respirations proceed simultaneously in T. selenatis. For bioremedial purposes, the presence of nitrate is desirable, since nitrate respiration induces the nitrite reductase that catalyzes selenate reduction to elemental selenium (DeMoll-Decker & Macy 1993). SES-3, in contrast, prefers nitrate as an electron acceptor and will not respire selenate until nitrate is depleted (Steinberg et al. 1992). The agricultural drainage waters contain about 50 mg/L nitrate, 150 times the selenate concentration, making this an important consideration (Macy et al. 1993). As the mobile, highly toxic Se(VI) is the dominant species found in the Central Valley’s San Joaquin River and the various evaporation ponds (Oremland et al. 1989; Thompson-Eagle et al. 1989), microbial reduction has the potential to remove selenate from soil porewaters and immobilize it as Se(0) an insoluble, much less bioavailable form. Studies of selenate reduction in evaporation pond surface sediment slurries have shown that anaerobic conditions, indigenous microorganisms are able to remove up to 99.7%

Table 2. Free energies of electron acceptor reduction coupled to H2 oxidation Reaction

DG* (kcal/mol e))

1/4 O(g) + 1/2 H2 fi 1/2 H2O 1/5NO3) + 1/5H+ +1/2H2 fi 1/10N2(g) + 1/5H2O 1/2MnO2(s) + H+ + 1/2 H2 fi 1/2Mn2+ + H2O 1/2UO22+ + 1/2 H2 fi H+ + 1/2UO2(s) 1/2SeO42) +1/2H+ + 1/2H2 fi 1/2HSeO3) + 1/2H2O 1/2 CrO42) + 5/3H+ + 1/2 H2 fi 3/2Cr3) + 4/3H2O Fe(OH)3(s) + 2H+ + 1/2H2 fi Fe2+ + 3H2O 1/4HSeO3) + 1/4H+ + 1/2 H2 fi 1/4Se + 3/4H2O 1/2H2AsO4) + 1/2H+ + 1/2H2 fi 1/2H2AsO3 + 1/2H2O 1/4SO42)+1/4H++1/2 H2 fi 1/4 HS)+1/2H2O

)23.55a )20.66a )22.48b )18.89c )15.53b )10.76b )10.49b )8.93b )5.51d )0.10a

PH2 = 1066 atm; PN2 = 0.78 atm; PO2 = 0.21 atm; pH = 7.0. 2+ 3+ ) 2

2

2
]=[UO2þ ]=[Fe2+]=[HSeO
[NO
2 ]=[Mn 3 ]=[H2 AsO4 ] =[H2AsO3]=[SO4 ]=[HS ]= 2 ]=[SeO4 ]=[HSeO3 ] =[CrO4 ]=[Cr [10)6] M. a Calculated from Zehnder & Stumm 1988. b Calculated from Morel & Hering 1993. c Calculated from Rusin et al. 1994. d Calculated from Vanysek 1995.

133 of the added selenate (approx. 300 lM) from porewater within 7 days of incubation, with addition of lactate, acetate and H2 stimulating reduction rates (Oremland et al. 1989). In pond sediments, selenate reduction occurs primarily near the surface, in the same region as denitrification and vertically above sulfate reduction. Areal rates of selenate removal from pond sediment porewaters have been estimated at up to 300lmol/m2/day (Figure 18). Bioremediation of selenate contamination by microbial reduction and precipitation has attracted many advocates; nevertheless, the longterm stability of elemental selenium in oxic soils is questioned (Frankenberger Jr. and Karlson, 1994b). Because selenite and selenate are soluble and relatively mobile, aqueous extraction from soils may be a better alternative. Microbial reduction might then be used to recover selenium from the aqueous phase. This technique has been extremely successful in various wastewaters, consistently removing 95–98% of the Se (Lawson and Macy, 1995; Macy et al. 1993) 3.3.2. Uranium reduction Uranium contamination of soils and groundwaters has resulted from the release of wastes from mining, extraction from ores, nuclear fuel reprocessing and ammunitions manufacture. Uranium

exists in most oxygenated wastes as the U(VI) uranyl ion UO22+, in reducing environments such as groundwater (Macaskie, 1991). While U(VI) is soluble and immobile, U(IV) is mobile, and highly insoluble in most aqueous solutions (Lovley & Phillips 1992a, b). Several years ago, the iron-respiring microorganisms Strain GS-15 and Alteromonas purefaciens (later renamed Geobacter metallireducens and Shewanella purefaciens, respectively) were shown to respire uranium as well as iron. The group of uranium-reducing microorganisms now includes several sulfate-reducing bacteria, and mechanistic investigation of Desulfovibrio vulgaris has shown that the respiratory electron transport protein cytochrome c, catalyzes the U(VI) reduction (Lovley et al. 1993b). Sulfate-reducing bacteria do not appear to obtain energy for growth from the process, however (Lovley et al. 1993a). Microbial uranium reduction, clearly applicable to precipitation of uranium from wastewaters, has also found a role in uranium removal from soils. Because bicarbonate ions form very strong, soluble complexes with U(VI), investigators used a 100 mM solution to extract uranium from contaminated soils, uranium ores and mill tailings. The bicarbonate was not able to remove the uranium completely (20–94% recovery) although the extracted uranium might have been

Figure 18. Rates of reduction and porewater concentrations of sulfate, nitrate and selenate in an evaporation pond core from the San Joaquin Valley. (a) Sulfate reduction; (b) denitrification; (c) selenate reduction. Values indicate the mean of 3 samples (Reprinted from Oremland et al. (1990), with permission. Copyright 1990 American Chemical Society).

134 the most dangerous fraction, as unrelated studies have shown the bicarbonate removes a much higher percentage of soil uranium than is bioavailable (Sheppard & Evenden 1992). To this extract, kept anaerobic under N2 ACO2 (80:20), H2 was added as an electron donor and Desulfovibrio desulfuricansas as the catalyst. The cells reduced and precipitated 75–100% of the uranium from the leachate within 24 h, yielding a very pure, highly concentrated uraninite solid. The reduction was somewhat inhibited by dissolved organic matter in the extract, as evidenced by a yellow color, but peroxide addition to oxidize the organic matter eliminated the color and the inhibition (Phillips et al. 1995). A remarkable feature of this microbial U(VI) precipitation is that the Desulfovibrio cells can be freeze-dried and stored under air for up to six months without loss of uranium-reducing activity (Lovley & Phillips 1992b). 3.3.3. Plutonium reduction Plutonium, obtained from irradiation of uranium and assembled into nuclear weapons, is an extensive soil contaminant near the production sites at Hanford, WA and Rocky Flats, CO (Zorpette 1996). Plutonium is also a product of nuclear energy generation, although much of it can be reprocessed (Macaskie 1991). Like uranium, plutonium is a redox-active metal, with oxidation states of III to VI known to exist in aqueous solution. Oxides and hydroxides of Pu(IV), with extremely low solubilities estimated at ksp=1056.8 to 1057.8, appear to predominate in contaminated soils (Rusin et al. 1994). Pu (III) hydroxide, while still quite insoluble at Ksp=1022.6, is more soluble than Pu(IV) compounds. Because the reduction of Pu(IV), as hydrous PuO2(s) to Pu3+ is theorectically capable of supporting bacterial growth, with a reduction potential similar to that of Fe(III), researchers have investigated the potential for plutonium reduction and dissolution by iron-reducing Bacillus strains. Experiments showed that cultures of B. circulans and B. polymyxa did indeed dissolve solid plutonium, amounting to 85–91% of the hydrous PuO2(s) added, or 1.25 lmol Pu, in the presence of 50 mM nitrilotriacetic acid (NTA). Reduction was the most likely means of dissolution, although this was not conclusively demonstrated: plutonium-dependent growth of the

bacteria, indicating energy gain from the reduction, was also not shown. With further development, including the use of biodegradable complexing agents in place of NTA, microbial reduction could provide a new pathway for insitu plutonium remediation (Rusin et al. 1994).

3.3.4. Arsenic reduction Microbial arsenic reduction, like plutonium reduction, transforms a relatively immobile species into a more mobile one and thus has the potential to operate in conjunction with soil washing to mobilize soil-bound arsenic. Inorganic arsenic in neutral aqueous solution occurs as As(V), arsenate (H2AsO4), and As(III), arsenite (H3AsO3). As(-III), occurring as arsine (AsH3), is a volatile reactive species formed by certain fungi and bacteria (Cheng & Focht 1979). Arsenate adsorbs strongly to iron and manganese oxyhydroxides and is thus very difficult to remove from soils and sediments. Arsenite, although more toxic, is more mobile, suggesting that reduction may promote arsenic mobilization (Bodek et al. 1988; Kuhn & Sigg 1993). Recently, several species of arsenate-respiring bacteria have been isolated from arsenic-rich substrates, including Strain MIT-13 (renamed Geospirillum arsenophius) (Ahmann et al. 1994). Its close relatives that also reduce selenate, are Strain SES-3 (Laverman et al. 1995) Desulfotomaculum auripigmentum (Newman et al. 1997) and Chrysiogenes arsenatis (Macy et al. 1996). Arsenic reduction is rapid in these microorganisms, measured at 2–5 mmol As-ml culture1 Æ day1 in both G. arsenophilus (Ahmann et al. 1994) and in Strain SES-3 (Laverman et al. 1995). The respiratory action of G. arsenophilus is able to dissolve arsenic from contaminated sediments (approximately 1000 ppm As) and other arsenical solids (Ahmann et al. 1997), and continuous-flow pilot-scale studies are in progress to maximize the arsenic-mobilizing potentials of the microorganism and others indigenous to contaminated soils.

3.3.5. Chromium reduction Cr(III) found at trace concentrations of 4–90 ppm in pristine areas, is the most stable and abundant form of chromium in most soils.

135 Kinetically it is quite inert, forming long-lived aqueous complexes with ammonia, sulfates, halides and organic acids. At slightly acidic to alkaline pH, however, ionic Cr(III) species tend to precipitate as amorphous Cr(OH3) if Fe3+ is present, with the result that much of the Cr(III) in soils is insoluble and immobile (Krishnamurthy & Wilkens 1994). Although Cr(III) is an essential trace nutrient, it does not readily enter living cells. Industrially-produced and -discharged Cr(VI), in contrast occurs as soluble, highly toxic and carcinogenic chromate and dichromate. Hexavalent chromium is freely taken up by cells and reduced intracellularly to Cr(III), which appears to mutagenize. Because Cr(III) is significantlly less toxic and less mobile than Cr(VI)-contaminated soil, numerous studies have focused on reducing Cr(VI) in-situ, by chemical (e.g., ferrous iron) or biological means. The intent is thus to ‘‘fix’’ the chromium in the soil in a harmless form. This approach requires careful management, since reoxidation of Cr(III) to Cr(VI) is catalyzed by manganese oxides and by molecular oxygen. The oxidation appears to be quite slow, primarily limited by solubility and mobility of the Cr(III)-laden soil is still quite possible under oxidizing conditions. The two essential requirements for safe chromium remediation by reduction, therefore, are (1) a permanent reducing environment and (2) immobilization, by precipitation or by complexation with degradation-resistant organic polymers, of the reduced chromium (Bartlett 1991; Krishnamurthy & Wilkens 1994; Losi et al. 1994a). Bioremediation efforts to date have focused on the study of Cr(VI)-reducing microorganisms. Numerous such bacteria have been isolated, including species of Bacillus, Desulfovibrio, Achromobacter, Aeromonas, Escherichia, Enterobacter and Pseudomonas (Lovley & Phillips 1994; Turick et al. 1996). Mechanisms of chromate reduction in anaerobic bacteria appear to involve respiratory electron-transport enzymes (Bopp & Ehrlich 1988; Shen & Wang 1993), with convincing evidence for catalysis by cytochrome c3 in Desulfovibrio vulgaris (Lovley & Phillips 1994). While it is thermodynamically possible for chromate reduction to generate enough energy to support respiration (Table 2), true chromate respiration remains to be discovered (Lovley & Phillips 1994; Shen & Wang 1993; Turick et al.

1996). In aerobic chromate-reducers, reduction may be fortuitously catalyzed by enzymes that have other natural substrates (Cervantes 1991). The presence of reactive electron donors appears to be the single greatest factor promoting chromate reduction both chemical and biological, in soils. In one illustrative set of experiments, sterile soils alone reduced almost half of the 800 lg/kg Cr(VI) present within 8 days. Addition of organic matter (cattle manure) improved the reduction to approximately 63%, and native microbial activity in organic-amended samples resulted in over 96% reduction, emphasizing the effectiveness of biological catalysis (Figure 19). Studies simulating field conditions have supported these results, confirming (1) the ubiquity of indigenous Cr-reducing microorganisms in a variety of soil, both contaminated and clean, (2) the importance of irrigation to maintain reducing conditions and (3) the proportionality between organic matter loading and Cr(VI) reduction (Cifuentes et al. 1996; Losi et al. 1994a, b), suggesting that Cr(VI) bioremediation by reduction in soil holds great promise in cases where reoxidation can be permanently prevented. 3.4. Volatilization 3.4.1. Selenium volatilization Microbial transformation of selenium into volatile methylated species, principally dimethylselenide (CH3)2Se with smaller amounts of dimethyldiselenide (CH3)2(Se)2 and dimethylselenone (CH3)2 (SeO2), removes selenium from soil and disperses it into the atmosphere. Airborne dimethylselenide (DMSe) reacts with hydroxyl radicals and ozone to yield oxidized products which are then thought to become associated with particles or aerosols that have residence times of several days, during which they may be dispersed over great distances (Frankenberger Jr. & Karlson 1994a). Microbial volatilization was first recognized as an important part of the selenium biogeochemical cycle in 1987, and it is now estimated that soils and plants emit 1700– 3300 metric tons of selenium into the atmosphere in the Northern hemisphere each year (Cooke & Bruland 1987; Frankenberger Jr. & Karlson 1995). Selenium methylation produces substances of greatly reduced toxicity and is thus thought to

136

Figure 19. Reduction of Cr(VI) over time in moist soils under sterile and non-sterile conditions, amended with cattle manure (OM) at 0 and 50 tons ha)1. Final values followed by different letters are statistically different (Reprinted from Losi et al. 1994a,b, with permission. Copyright 1994 Society for Environmental Toxicology and Chemistry).

protect the methylators: dissolved dimethylselenide is 500–700 times less toxic than selenate and selenite, while inhaled dimethylselenide is non-toxic in concentrations at up to 8034 ppm. Selenium volatilization activity appears to be widespread among aerobic microorganisms, including the native microbiota of the Kesterson evaporation ponds and additional fungi of the genera Alternaria, Scopulariopsis, Fusarium, Acremonium, Pencillium and Ulocladium (Frankenberger Jr. & Karlson 1995). The biomethylation pathway has not been fully elucidated, but is thought to require reduction of selenium species to Se2), followed by methylation; as a result, selenite (Se(IV)) is volatilized more rapidly than is selenate (Se(VI)) (Frankenberger Jr. & Karson 1994a). Selenium volatilization is controlled both by enzymatic activities and vapor pressures of the methylated gases, with the result that the process is highly temperature-dependent (Q10=1.7 from 4–53 C). The process thus shows pronounced seasonal fluctuations, with average emission rates of 75–100 lg Se/h/m2 in the summer and 10–50 lg Se/h/m2 in the winter in managed dewatered evaporation ponds containing only indigenous microorganisms. Other factors were also important in these pilot study plots; soil moisture contents of at least 70% optimized Se

emissions, although saturation, which could have created anoxia or leached selenium into deeper soil layers, was avoided. Carbon addition in the form of cattle manure also increased volatilization rates and frequent tillage improved contact between air and the soil solution, presumably releasing methylated compounds and stimulating aerobic metabolism. Nearly 60% of the selenium in the soils (mean original content: 11.4 ppm) was removed in 22 months, and 18-fold enhancement over removal from untreated plots. The time needed to reach the cleanup goal of 4 ppm was estimated to be 2.6 years (Frankenberger Jr. & Karlson 1994a, 1995). Microbial volatilization of selenium is a promising method of bioremediation and one which may be further improved when the relevant molecular mechanisms are understood. An important limitation, however, is the microbial ability to metabolize only water-soluble selenium, apparently because only these forms may be taken up by microorganisms (Frankenberger Jr. & Karlson 1994a). While this pool may represent the most hazardous selenium present, significant amounts of soil-bound but labile selenium may resist volatilization, necessitating long treatment periods or the assistance of solubilizing agents to reach cleanup standards.

137 3.4.2. Mercury volatilization Mercury resistance systems have been found in all bacterial groups tested and consistently include an operon of six genes: merR, a regulatory gene: merD, a second regulator, merP and merT, encoding proteins that transport mercuric ions into the cell; merA, encoding a mercuric reductase that transforms Hg2+ to volatile Hg0 and merB, encoding an organomercurial lyase that cleaves Hg from organomercury compounds (Figure 20). The bioremedial potential of this system lies in the ability of the bacteria to remove highly toxic mercuric ions from a substrate, releasing them as elemental mercury of greatly reduced toxicity that can then be trapped or dispersed into the atmosphere. The system is quite well understood and although it is ubiquitous among bacteria, it is showing greatest promise within transgenic plants for mercury removal from soils (Rugh et al. 1996; Silver 1994).

4. Metal-microbe interactions and their impact on bioremediation of metal contaminated soils and sediments Estimates of the global market for the cleanup and prevention of metal contamination vary, but conservative calculations suggest that the current market for metal bioremediation may rise to $200 billion in the US alone by 2005. The emerging market for the clean up of radioactive contamination in the US may already be worth as much as $300 billion (McCullough et al. 1999).

Unfortunately, existing chemical techniques are not always economical or cost effective for the remediation of water or land contaminated with metals and radionuclides. Current strategies for land contaminated with metals include the use of ‘‘dig and dump’’ approaches that only move the problem to another site, are expensive and impractical for large volumes of soil or sediment. Likewise soil washing, which removes the smallest particles that bind most of the metals, is useful but can be prohibitively expensive for some sites. ‘‘Pump and treat’’ technologies rely on the removal of metals from the site in an aqueous phase which is treated ex situ (e.g. above land). These approaches can cut down on excavation costs but are still expensive, and metal removal can be inefficient. A potentially economical alternative is to develop biotechnological approaches that could be used in the sediment or soil (in situ) to either extract the metals or stabilise them in forms that are immobile or non-toxic. In addition to developing biotechnologies to treat contaminated land, there is also considerable interest in more effective techniques that can be used to treat water contaminated with metals from a range of industrial processes. Problems inherent in currently used chemical approaches include a lack of specificity associated with some ion exchange resins, or the generation of large quantities of poorly settling sludge through treatment with alkali or flocculating agents. This section gives an overview of metal–microbe interactions, and describes how they could be harnessed to clean up metal contaminated, soil and sediments.

Figure 20. Genes and polypeptides of mercury resistance systems from Gram-negative Serratia sp. Aa, length of gene products in amino acids; P/O, promoter/operator region (Adapted from Ji & Silver 1995, with permission of the Society for Industrial Microbiology).

138 4.1. Metal and radionuclide contamination The US EPA (http://www.epa.gov) lists four main groups of priority compounds which include polychlorinated biphenyls (PCBs), base neutrals and acids (e.g. phenol and naphthalene) and volatile organic compounds (VOCs), including carbon tetrachloride and BTEX contaminants. Trace metals that are recoverable and dissolved are also listed; these include antimony, chromium, mercury, silver, arsenic, copper, nickel, thallium, beryllium, lead, selenium, zinc and cadmium. In addition to these toxic metals, radionuclides are priority contaminants at nuclear installations worldwide, and will require treatment. For example, in the US Department of Energy’s 120 sites contains 1.7 trillion gallons of contaminated ground water and 40 million cubic metres of contaminated soil and debris. More than 50% of the sites are contaminated with radionuclides, with the priority radionuclides being cesium-137, plutonium-239, strontium-90, technetium-99 and uranium-238 and uranium-235, in addition to toxic heavy metals including chromium, lead and mercury (McCullough et al. 1999). The US DOE Natural and Accelerated Bioremediation (NABIR) program is notable in its support of emerging biotechnological approaches for dealing with this legacy waste. In Europe, the EU Water Framework Directive (2000/60/EC) is the principal driver for environmental legislation, and also recognizes metals amongst the 32 substances on the priority list. These include lead, nickel, mercury cadmium and tin and their compounds. 4.2. Biosorption The term biosorption is used to describe the metabolism-independent sorption of heavy metals and radionuclides to biomass. It encompasses both adsorption; the accumulation of substances at a surface or interface and absorption; the almost uniform penetration of atoms or molecules of one phase forming a solution with a second phase (Gadd & White 1989). Both living and dead biomass are capable of biosorption and ligands involved in metal binding include carboxyl, amine, hydroxyl, phosphate and sulfhydryl groups. Biosorption of metals has been reviewed extensively (McHale & McHale

1994; Tobin et al. 1994; Volesky & Holan 1995; Beveridge et al. 1997a,b; Lloyd & Macaskie 2000). Biosorption is generally rapid and unaffected over modest temperature ranges and in many cases can be described by isotherm models such as the Langmuir and Freundlich isotherms (Volesky & Holan 1995). Gadd and White (1989), however, noted that more complex interactions are difficult to model because the adsorption of solutes by solids is affected by factors including diffusion, heterogeneity of the surface and pH. An additional isotherm, the Brunauer– Emmett–Teller (BET) isotherm, which assumes multilayer binding at constant energy has also been used to describe metal biosorption (de Rome & Gold 1991; Andres et al. 1993). This model assumes that one layer need not necessarily be completely filled before another is commenced. Further insight is offered by Andres et al. (1993), who summarize that each adsorption layer of the BET model can be reduced to Langmuir behavior with homogeneous surface energy, in contrast to the adsorption energy requirements of the Freundlich isotherm. Other studies have used complex multistage kinetic approaches to model biosorption (Weidemann et al. 1981; Treen-Sears et al. 1984). Ultimately, however, the amount of residual metal remaining in solution at equilibrium is governed by the stability constant of the metal– ligand complex (Macaskie 1991), and the only way to change the equilibrium position is to modify the binding ligand to one which has a greater binding affinity for the given metal, or to transform the metal from a poorly sorbing species to one which has a higher ligand-binding affinity, e.g. by a change of metal valence. Alternatively some new studies have applied the tools of molecular biology to enhance metal sorption and do warrant attention. For example, a mouse metallothionein was targeted to the outer membrane of a metal-resistant Ralstonia eutropha isolate (Valls et al. 2000). The engineered strain accumulated more Cd2+ than its wild type counterpart, and offered tobacco plants some protection from Cd2+ when inoculated into contaminated soil (Valls et al. 2000). A surface display technique has also been used successfully to generate ZnO binding peptides fused to fimbrae on the surface of cells of Escherichia coli (Kjaergaard et al. 2000). Finally, Gram-positive

139 bacteria (Staphylococci) have also been engineered to produce surface exposed peptides, able to bind Hg2+ and Cd2+ (Samuelson et al. 2000). Enhanced uptake of cadmium and mercury by Escherichia coli expressing a metal-binding motif has also been reported (Pazirandeh et al. 1998). Future studies could usefully compare the performance of such engineered systems with other more traditional biosorbants. However, it has been noted that despite the relatively long time during which biosorption has been studied, and considerable early interest, there has been reduced interest in commercialising this type of technology, and the authors are aware of no current commercial applications of this type of technology. 4.3. Metabolism-dependent bioaccumulation Energy-dependent metal uptake has been demonstrated for most physiologically important metal ions, and some toxic metals and radionuclides enter the cell as chemical ‘‘surrogates’’ using these transport systems. Monovalent cation transport, for example K+ uptake, is linked to the plasma membrane-bound H+-ATPase via the membrane potential, and is, therefore, affected by factors that inhibit cell energy metabolism. These include the absence of substrate, anaerobiosis, incubation at low temperatures and the presence of respiratory inhibitors such as cyanide (White & Gadd 1987). The requirement for metabolically active cells may, therefore, limit the practical application of this mode of metal uptake to the treatment of metals and radionuclides with low toxicity and radioactivity. For example, in the study of White and Gadd (1997), increasing metal concentrations inhibited H+ pumping, potentially deenergizing the cell membrane and reducing cation uptake. Although the presence of multiple transport mechanisms of differing affinities may cause added complication, metal influx frequently conforms to Michaelis–Menten kinetics (Borst-Pauwels 1981). Once in the cell, metals may be sequestered by cysteine-rich metallothioneins (Higham et al. 1984; Turner & Robinson 1995) or, in the case of fungi, compartmentalized into the vacuole (Gadd & White 1989; Okorov et al. 1977). In this context, it should be emphasized that the uptake of higher mass radionuclides, e.g. the actinides into

microbial cells has been reported sporadically and remains poorly characterised (Lloyd & Macaskie 2002; Lloyd & Macaskie 2000). Metabolism-dependent bioaccumulation of metals by microorganisms has not been used commercially for bioremediation purposes. 4.4. Enzymatically catalysed biotransformation Microorganisms can catalyze the transformation of toxic metals to less soluble or more volatile forms. For example, the microbial reduction of Cr(VI) to Cr(III), Se(VI) to Se(0), V(V) to V(III), Au(III) to Au(0), Pd(II) to Pd(0), U(VI) to U(IV) and Np(V) to Np(IV) results in metal precipitation under physiological conditions and has been reviewed in detail recently (Lloyd 2003). In many cases the high valence metal can be used as an electron acceptor under anoxic conditions. A good example here is the reduction of soluble U(VI) to insoluble U(VI) by Fe(III)-reducing bacteria (Lovley et al. 1991), which has been harnessed recently to remediate sediments contaminated with uranium in situ. Metal detoxification is also possible through biotransformations of toxic metals. The bioreduction of Hg(II) to relatively non-toxic Hg(0) is perhaps the best studied example. Finally, in addition to redox biotransformations, biomethylation may also increase the volatility of metals, with the methylation of mercury, cadmium, lead, tin, selenium and tellurium recorded (Diels et al. 1995). These mechanisms could also offer potential use for the in situ remediation of contaminated soil (Losi & Frankenberger 1997). As mentioned above, microbes have evolved sophisticated approaches to deal with toxic metal (Bruins et al. 2000), often involving redox transformations of the toxic metal. Perhaps the best studied metal resistance system is encoded by genes of the mer or mercury resistance operon (Hobman et al. 2000). Recent studies have confirmed the biotechnological potential of this widespread resistance determinant. Hg(II) is bound in the periplasm of Gram-negative bacteria by the MerP protein, transported into the cell via the MerT transporter, and detoxified by reduction to relatively non-toxic volatile elemental mercury by an intracellular mercuric reductase (MerA). Mer proteins are expressed under the regulation of the activator protein, MerR,

140 which binds Hg(II) and activates gene expression. Mercury-resistant bacteria and the proteins that they encode have been used recently for the bioremediation of Hg-contaminated water and in the development of biosensors for bioavailable concentrations of Hg(II) (Bontidean et al. 2002). Such sensors may prove very useful in identifying the need for metal remediation, which will be dictated in many cases by the concentration of bioavailable toxic metals in a given soil or water matrix, and also in defining the end point for bioremediation efforts. Focusing on Hg biosensors, several different components of the mer system have been used in prototype sensors, including the NADPH-dependent mercuric reductase (MerA) in an enzymelinked biosensor (Eccles et al. 1997), the mer regulatory region in a whole cell biosensor (Geiselhart et al. 1991), and the MerR protein in a capacitance biosensor (Bontidean et al. 2000). 4.5. Biomineralization via microbially generated ligands Microorganisms are able to precipitate metals and radionuclides as carbonates and hydroxides via plasmid-borne resistance mechanisms, whereby proton influx countercurrent (antiport) to metal efflux results in localized alkalinization at the cell surface (Diels et al. 1995; Van Roy et al. 1997). Alternatively, metals can precipitate with enzymatically generated ligands, e.g. sulfide (Barnes et al. 1991), or phosphate (Macaskie et al. 1992). The concentration of residual free metal at equilibrium is governed by the solubility product of the metal complex (e.g. 10)20 to 10)30 for the sulfides and phosphates, higher for the carbonates). Most of the metal should be removed from solution if an excess of ligand is supplied. This is difficult to achieve using chemical precipitation methods in dilute solutions; the advantage of microbial ligand generation is that high concentrations of ligand are achieved in juxtaposition to the cell surface, that can also provide nucleation foci for the rapid onset of metal precipitation; effectively the metals are concentrated ‘‘uphill’’ against a concentration gradient. This was demonstrated using the gamma isotope 241 Am supplied at an input concentration of approximately 2.5 parts per billion; approximately 95% of the metal was removed as bio-

mass-bound phosphate (Macaskie et al. 1994), with the use of gamma-counting permitting detection at levels below those of most analytical methods. In many cases, the production of the ligand can also be fine-tuned by the application of Michaelis–Menten kinetics. Sulfide precipitation, catalysed by a mixed culture of sulfate-reducing bacteria, has been utilized first to treat water co-contaminated by sulfate and zinc (Barnes et al. 1991) and also, soil leachate contaminated with sulfate alongside metal and radionuclides (Kearney et al. 1996). Ethanol was used as the electron donor for the reduction of sulfate to sulfide in both examples. A later study has confirmed the potential of integrating the action of sulfur cycling bacteria (White et al. 1998). In the first step of a twostage process, sulfur-oxidising bacteria were used to leach metals from contaminated soil via the generation of sulfuric acid, and in the second step the metals were stripped from solution in an anaerobic bioreactor containing sulfate-reducing bacteria (Kearney et al. 1996). The ubiquitous distribution of sulfate-reducing bacteria in acid, neutral and alkali environments (Postgate 1979), suggests that they have the potential to treat a variety of effluents, while the ability of the organisms to metabolize a wide range of electron donors, may also allow co-treatment of other organic contaminants. Recent work by Paques BV of the Netherlands have confirmed the potential of sulfate-reducing bacteria to treat metal waste in ex situ bioreactors for the treatment of a wide range of metal-contaminated waters. Bioprecipitation of metal phosphates via hydrolysis of stored polyphosphate by Acinetobacter spp. is dependent upon alternating aerobic (polyphosphate synthesis) and anaerobic (polyphosphate hydrolysis and phosphate release) periods (Boswell et al. 2001). This obligately aerobic organism is fairly restricted in the range of carbon sources utilized, but the preferred substrates (acetate and ethanol) are widely and cheaply available. The best-documented organism for metal phosphate biomineralization, a Citrobacter sp., (now reclassified as a Serratia sp. on the basis of molecular methods, the presence of the phoN phosphatase gene and the production of pink pigment under some conditions (Pattanapipitpaisal et al. 2002) grows well on cheaply available substrates and viable cells are

141 not required for metal uptake since this relies on hydrolytic cleavage of a supplied organic phosphate donor (Macaskie et al. 1992). The expense of adding organic phosphate was calculated to be the single factor which limited the economic viability of this approach (Roig et al. 1995); moreover organophosphorus compounds are often highly toxic. A possible alternative phosphate donor, tributyl phosphate (TBP), is used widely as a cheap solvent and plasticizer but its degradation is more difficult because this compound is a phosphate triester, with three cleavage events per molecule required to liberate one molecule of phosphate. TBP is biodegradable; this activity was unstable (Thomas & Macaskie 1998) but biogenic phosphate from TBP hydrolysis was harnessed to the removal of uranium from solution in a flow-through system (Thomas & Macaskie 1996). The enzymatic activity responsible for TBP hydrolysis remains obscure but a recent report of TBP hydrolysis by enzymatic degradation was studied using cell free extracts of Serratia odorifera (Berne 2004). As an alternative to TBP and glycerol 2-phosphate the possibility exists for the harnessing of phytic acid degrading organisms to metal accumulation. Phytic acid, widely and cheaply available as a natural plant product, accounts for up to 50% of the total organic P in soils (Rodriguez & Fraga 1999). The additional advantage of phytic acid is that it contains 6 phosphates/molecule (c.f. 1 phosphate/molecule in TBP and G2P). Several phytases have been cloned and characterized, including one from the related organism E. coli (Greiner et al. 1993) with the phytase reaction mechanism described. Preliminary studies have suggested that E. coli phytase can be applied to metal removal (Patersen–Beedle et al. unpublished).

target metal precipitation reactions (Macaskie et al. 1996). The priming deposit is made initially by the sulfide or phosphate biomineralization routes described above. With addition of Fe as the precipitant metal to sulfate-reducing bacteria producing H2S, cellbound FeS then acts as a sorbent for ‘target’ metals (Ellwood et al. 1992; Watson & Ellwood 1994, 1988). The use of FeS is notable because this provides the mechanism for rapid biomass separation from the liquor via the magnetic properties of Fe in a high-gradient magnetic separator (Ellwood et al. 1992; Watson & Ellwood 1994; Watson & Ellwood 1988). Metal phosphate can also be used as the priming deposit, in two ways. LaPO4 pre-deposited onto metal-accumulating Citrobacter sp. via biogenic phosphate release can be used as the ‘priming’ deposit for subsequent deposition of actinide phosphates (Macaskie et al. 1994). Also, by use of a priming deposit with an appropriate, well-defined cystalline lattice, the target metal can be intercalated within the ‘host’ lattice, effectively by a mechanism of bioinorganic ion-exchange, well-established as a chemical process (Clearfield 1988; Pham-Thi & Columban 1985; Pozas-Tormo et al. 1987) and best-characterized biologically in the case of Ni2+ removal into biomass-bound HUO2PO4ÆnH2O (Bonthrone et al. 1996). Within the context of nuclear waste remediation, the approach of MECHM, using biogenic hydrogen uranyl phosphate as the ‘host’ matrix, and using either intercalative ion exchange or co-crystallisation, has been applied to the removal of Cs, Co and Sr from ‘surrogate’ solutions and the isotopes 137Cs, 60Co and 90Sr from real nuclear wastes in South Korea (Paterson–Beedle et al. unpublished).

4.6. Microbially enhanced chemisorption of heavy metals

4.7. Metal remediation through biodegradation of associated organic compounds

Microbially enhanced chemisorption of heavy metals (MECHM) is a generic term to describe a class of reactions whereby microbial cells first precipitate a biomineral of one metal (‘priming deposit’). The priming deposit then acts as a nucleation focus, or ‘host crystal’ for the subsequent deposition of the metal of interest (‘target’ metal), acting to promote and accelerate

Citrate has found use as a chelating agent in decontamination operations forming highly soluble metal–citrate complexes, which can be degraded by microorganisms resulting in subsequent re-precipitation of metals (Francis 1994). Early work suggested that the type of complex formed between the metal and citric acid plays an important role in determining its

142 biodegradability, with the binuclear uranium–citrate complex being recalcitrant to microbial degradation (Gadd & White 1989), but subsequent photodegradation was possible for these problematic complexes. This opened up the way for a multi-stage process for the biodegradation of mixed metal–citrate complexes via microbiological and photochemical steps (Francis 1994). Cultures of Pseudomonas aeruginosa and P. putida were also able to grow using a range of metal– citrate complexes as carbon, with metal precipitation promoted by the addition of inorganic phosphate (Thomas et al. 2000). A unique aspect of this study was the use of P. putida to treat Ni– citrate waste generated by cleaning a bioinorganic ion exchange column previously used to remove Ni. Another industrial chelating agent ethylenediaminetetraacetate (EDTA) also forms very strong complexes with di- and trivalent metals that are degraded by a mixed microbial population (Thomas et al. 1998) and by bacterial strain DSM 9103 (Satroutdinov et al. 2000). The biodegradation of toxic organotin compounds used as biocides and antifouling agents has also received recent attention (e.g. the degradation of tributyl tin to di- and monobutyl tin reviewed in Gadd (2000)). 4.8. In-situ versus ex-situ remediation Although most of the processes described thus far have been developed primarily for ex situ treatment of contaminated water in situ remediation offers the potential for low cost treatment of groundwater contamination with metals. Ex situ remediation strategies such as pump and treat systems remain a method of choice for metal-contaminated sites and can be effective at removing significant quantities of contaminant metals from heavily contaminated (and highly concentrated) areas. For low contaminant concentration scenarios in situ immobilization of contaminants has been proposed as an attractive potential treatment option (Lovley et al. 1991; Abdelouas et al. 1998a, b, 1999; Finneran et al. 2002). In situ immobilization of contaminant metals relies on the stimulation of anaerobic microbial communities within the subsurface. Anaerobic microorganisms enzymatically reduce contaminant metals such as uranium, chromium, techne-

tium, vanadium and metalloids such as selenium thereby converting these metals to less soluble forms (Lloyd et al. 2000; Lovley & Coates 2000; Anderson & Lovley 2002; Carpentier et al. 2003; Ortiz-Bernard et al. 2004). Stimulation of metal reduction within the subsurface offers a relatively simple method to remove metal contaminants from groundwater, thereby limiting or greatly reducing further transport of contaminants downgradient (Lovley et al. 1991; Lovley 1993; Anderson & Lovley 2002). Efforts to understand and control in situ metal reduction are currently active areas of research within the U.S. Department of Energy (U.S. Department of Energy, O.o.B.a.E.R 2004) particularly in light of the recent focus on the potential ecological impacts of ex situ remediation practice (Whicker et al. 2004). A key issue in the development of in situ bioremediation systems for metal contaminants is the identification of target groups of microorganisms within the subsurface whose known physiology has the potential to affect contaminant metal mobility. Both Fe(III)- and sulfate-reducing organisms are known to enzymatically reduce contaminant metals such as U(VI), Cr(VI), Co(III) and Tc(VII) in laboratory cultures (Lovley et al. 1991; Lovley & Phillips 1992a, b; Gorby & Lovley 1992; Lovley 1993; Caccavo Jr. et al. 1994; Lloyd & Macaskie 1996; Gorby & Bolton 1998; Tebo & Obraztsova 1998; Lloyd et al. 2000) but for remediation purposes it is necessary to demonstrate that the appropriate microorganisms are present within the contaminated subsurface. Members of the Geobacteraceae (d-proteobacteria) are of particular note as these organisms have been identified as a dominant group within sediments upon the stimulation of Fe(III)-reducing conditions (Snoeyenbos-West et al. 2000; Finneran et al. 2002; Holmes et al. 2002; Anderson et al. 2003) and in contaminated sediment where Fe(III) reduction is a dominant process (Anderson et al. 1998; Ro¨ling et al. 2001; Cummings et al. 2003). Geobacteraceae were also recently detected within the groundwater of an uraniumcontaminated aquifer during a test of stimulated in situ U(VI) reduction (Anderson et al. 2003). Geobacteraceae, while best known as dissimilatory Fe(III)-reducing bacteria, also couple the oxidation of simple organics to the reduction of contaminant metals such as uranium, chromium, technetium and vanadium (Chen & Hao 1998;

143 Lloyd et al. 2000; Anderson & Lovley 2002; Oritz-Bernad et al. 2004). Additionally, the product of dissimilatory Fe(III) reduction, Fe(II), is a potential abiotic reductant for contaminant metals such as Tc(VII), V(V) and Cr(VI) (Fendorf et al. 2000; Lloyd et al. 2000; Oritz-Bernad et al. 2004). The demonstrated enrichment of Geobacteraceae within an aquifer during an in situ trial of stimulated uranium reduction and the potential abiotic benefits of in situ Fe(II) production suggest Geobacteraceae to be a target group of microorganisms within the subsurface for removing metals from groundwater. Further investigation of stimulated subsurface metal reduction will likely reveal other groups of bacteria, such as sulfate reducers, that could also play an important role in contaminant metal reduction and stabilization within subsurface environments.

5. Role of biofilms in bioremediation of metal polluted soils and sediments 5.1. Biofilms in natural environment Bacteria are widespread in soil and aquatic environments and are predominantly found in biofilm communities. A biofilm may be defined as a gelatinous extracellular polymeric substance (EPS) of biological origin containing a consortium of bacteria, fungi, and algae. Biofilms are present at the solid–aqueous interface throughout the aquatic environment (Lappin-Scott et al. 1993) and typically at the sediment–water interface (Davison et al. 1997; Fo¨rstner, 2004). Biofilms form when bacterial consortia attach to mineral surfaces and produce films of hydrated extracellular polymers (Jackson et al. 1999). Once biofilms form, the bacterial-mineral micro-aggregates create complex interfaces with the surrounding aqueous solution. The biofilms may act as an ‘‘insulating layer’’ between the solution and the mineral surface or form ‘‘micro-environments’’ in which the local solution conditions are different than those in the bulk solution (Glazer et al. 2002). Reactive functional groups, such as carboxyl, hydroxyl, amino and phosphoryl moieties, present on the bacterial surfaces and exopolysaccharide matrix, can provide a large array of binding sites for metals (Flemming & Leis 2002). For instance, autotrophic biofilms have

been observed to sequester metals at 4–5 orders of magnitude above the concentration in the surrounding waters (Liehr et al. 1994). Biofilms may simultaneously block surface sites on the underlying substrate (Templeton et al. 2003a,b). In addition, bacterial activity may catalyze the trans-formation of toxic metals into less (or more) toxic species or enhance the dissolution of the underlying mineral substrate and bulk solution surrounding of the biofilm (Beyenal et al. 2004; Labrenz & Banfield 2004). All of these combined interactions may strongly affect the mechanisms and kinetics of metal sorption reactions in soils and aquatic environments (Newman & Banfield 2002; Gadd 2004). However, biofilms that form on solid phases in natural contaminated environment sites may differ significantly in form and function from those found in other environments described to date. So far biofilms have been mainly used to remove inorganic contaminants from aqueous systems (industrial wastewaters) (Diels et al. 2003, Remoudaki et al. 2003; Wagner-Do¨bler 2003; van Hullebusch et al. 2003; Bender et al. 2004, Bender & Phillips 2004). However, they been also shown to influence the mobility of inorganic contaminants in subsurface environment such as soils and sediments. Ecological success of biofilms and their broad array of microbial activities suggest that these microbial ecosystems might be useful to bioremediation of environmental pollutants as described in the following subsection. 5.2. Biofilm and metals/radionuclides immobilization Only the main mechanisms involved in the immobilization of metal and radionuclides, as shown Figure 21, will be described in the following subsection. 5.2.1. Passive processes: biosorption of metals by biofilms Microorganisms adsorb and concentrate many metallic cations through electrostatic interactions with anionic carboxyl and phosphoryl groups in the cell wall. Biosorption consists of several mechanisms, mainly ion exchange, chelation, adsorption, and diffusion through cell walls and membranes. Biosorption encompasses those

144 physico-chemical mechanisms by which metal species and radionuclides are removed from an aqueous solution. This phenomenon is often attributed to the binding of metals onto the bacterial cell surface (White et al. 1995; Cox et al. 1999). In addition, however, the EPS that surround cells are also very reactive and can readily bind metals such as iron (Konhauser et al. 1993, 1994). A growing body of literature shows evidence that EPS bind and concentrate a range of inorganic contaminants. EPS is a highly hydrated material (99% water) possessing an extremely porous fibrillar structure that renders it highly adsorptive. The EPS consist of polysaccharides, proteins, lectins, lipids, nucleic acids, etc... The structure of the EPS-matrix is considerably stronger in the presence of different cations, which interact with exposed carboxyl groups on the EPS. EPS can bind a wide variety of metals, including Pb, Sr, Zn, Cd, Co, Cu, Mn, Mg, Fe, Ag, and Ni (Ferris et al. 1989). The sorption capacity of biofilms can be attributed to the complex or chelate formation of the EPS-matrix with different cations and the uptake of cations

by algae and bacteria living in the biofilms. The binding capacity of a given EPS-matrix also depends on its physical state (gel versus slime versus dissolved state) and the pH of the water. In order to reach stronger binding of the metal ions, the gel structure and higher pHs (basic conditions) are favourable. Ferris et al. (1989) reported that near neutral pH, microbial biofilms concentrated metals up to 12 orders of magnitude higher than observed under acidic conditions. Neutral to alkaline microenvironments are commonly produced within microbial mats and biofilms through such processes as photosynthesis and sulfate reduction (Krumbein 1979). 5.2.2. Microbial mineralization Minerals formed in association within biofilms are commonly referred to as biominerals, e.g. sulfate, phosphate, carbonate, sulfide or silicate ions (Schultze-Lam et al. 1996). Given the ability of EPS to concentrate various metals, it is not surprising that bacteria have been implicated in a wide variety of biomineralization processes. As a result of cellular metabolism, microorganisms

Figure 21. Biofilm–metal interactions which have an impact on bioremediation of metal contaminated soils and sediments (Modified from van Hullebusch et al. 2003).

145 alter the chemical microenvironment around the cell, modulating the pH, as well as the concentration of a variety of organic and inorganic solutes. This can induce the large-scale precipitation of authigenic minerals in natural environments (Geesey et al. 1989; Beveridge et al. 1997a,b; Douglas and Beveridge, 1998; Langley & Beveridge 1999). For instance, stromatolites, made of calcium carbonate precipitates, are intimately associated with biofilms that comprise an abundant mix of cyanobacteria, bacteria and diatoms (Douglas & Beveridge 1998; Arp et al. 2001). Sometimes nanomineral phases can even form directly in the bacterial cytoplasm as has been described for acanthite (Ag2S) (Klaus et al. 1999) or amorphous iron hydroxide (Fortin 2004). The tolerance of biofilms to high metal concentrations may also be due to their ability to precipitate insoluble metal salts outside the cells as sulfides, phosphates, carbonates, oxides or hydroxides (Bender et al. 1995; Douglas & Beveridge 1998). 5.2.3. Oxidation and reduction Inorganic contaminant entrapment and adsorption by iron and manganese oxides. A microbial biofilm consortium catalyses oxidation of Fe(II) in acidic and near neutral solutions to produce iron (Fe(III)) hydroxides (Brown et al. 1998b, 1999). Iron oxide precipitates associated with biofilm EPS were identified as ferrihydrite in both pH conditions but the precipitate morphology and minor elemental composition was pH dependent (Ferris et al. 1989). Investigation of biofilm consortia from groundwater by Transmission Electronic Microscopy (TEM) showed that the accumulation of iron on the cell surfaces was stepwise and that iron precipitates were found at the cell surfaces and throughout the biofilm matrix (Brown et al. 1998a). Bacterial cell walls and EPS are generally negatively charged, which allows positively charged metal ions to be adsorbed. These metal ions may than act as nucleation sites for the deposition of more iron from solution as shown by Brown et al. (1998a). Many investigations report that the sorptive capacities of Fe/Mn oxides for metals are increased in the presence of microorganisms (Haack & Warren 2003; Nelson et al. 2002). The iron oxide present in the biofilm can adsorb zinc, lead and cadmium in the biofilm (Baldi et al.

2001). Lack et al. (2002a, b) showed that the anaerobic bio-oxidation of 10 mM Fe(II) by Dechlorosoma species and precipitation of Fe(III) oxides by these microorganisms resulted in rapid adsorption and removal of 55 lM uranium and 81 lM cobalt from solution. Metal reductive precipitation. Bioreduction is generally recognized as an important metabolic process controlling the fate and transport of heavy metals and radionuclides, as well as the over all geochemistry of subsurface and sedimentary environments. Bioreduction of multivalent metals can convert dissolved, oxidized forms of multivalent heavy metals and radionuclides (Cr, Mo, Se, U, Pd, Tc, Au), such as U(VI) and Cr(VI), to reduced forms that readily precipitate from solution. The mobility of most of these elements in the environment is strongly dependent on their chemical oxidation state. Under oxidizing conditions, they occur as highly soluble and mobile ions, however, under reducing conditions they usually form insoluble phases (Abdelouas et al. 2000). A recent review gives a detailed understanding of some of these transformations at a molecular level (Lloyd 2003). For instance, molybdenum (VI) reduction by D. desulfuricans in the presence of sulfide resulted in the extracellular precipitation of the mineral molybdenite MoS2(S) (Tucker et al. 1997). Lloyd et al. (2001) show that SRB reduce Tc(VII) and this metal precipitates as TcO2 at the periphery of the cell. Microbial reduction of toxic Se(VI) and Se(IV) to the colloidal elemental selenium [Se(0)] can be used to remove Se from wastewater (Lloyd et al. 2001; Hockin & Gadd 2003). Elemental selenium has not been localized on the bacterial cells (Lloyd et al. 2001) but was released in the bulk solution. SRB biofilms have been shown to reduce Cr(VI) to Cr(III), which precipitates as insoluble hydroxides (Cr(III)) (Smith & Gadd 2000).

5.2.4. Microbial precipitation Sulfide precipitation. Sulfate-reducing bacteria (SRB) are the important physiological group for sulfide production. SRB are commonly used for the bioremediation of metal-contaminated soil or water. White et al. (1998), for instance, described an integrated microbial process for the

146 bioremediation of soil contaminated with toxic metals using microbially catalyzed reactions. In this process, bioleaching of Cd, Co, Cr, Cu, Mn, Ni, and Zn via sulfuric acid produced by sulfur-oxidizing bacteria was followed by precipitation of the leached metals as insoluble sulfides by the action of SRB. Microorganisms have been used in the remediation of metal pollution in very specific cases. The most successful has been the reduction and the eventual precipitation of soluble metal sulphates as insoluble sulfides in liquid wastes by the use of the microorganisms that are part of the sulfur cycle, especially sulfate-reducing bacteria that use sulfate as their ultimate electron acceptor to produce hydrogen sulfide, which binds with the metals to give metal sulfide. This procedure is used nowadays not only for the treatment of surface waters but also for massive cleaning up of underground water and is available in many formats, even as commercial large-scale treatment plants. Bioprecipitation of metals by microbiologically produced sulfide (SRB) is then an effective way to immobilise a wide range of metals, including Zn, Cu, Pb, Ni and Cd. The metals precipitate as highly insoluble sulfides (entrapment, nucleation and crystallisation of insoluble sulfides), e.g. ZnS, CdS, CuS, CoS, NiS and FeS (White & Gadd 1998; White & Gadd 2000; Labrenz et al. 2000; Wang et al. 2001; Drzyzga et al. 2002; Valls & de Lorenzo, 2002; Utgikar et al. 2002; White et al. 2003; Krumholz et al. 2003). Sulfide precipitation is an efficient means of removing toxic metals from solution, while immobilisation in a biofilm is also an affective method to reduce the mobility of the metals, which is of value in bioremediation of anoxic sediment and wetlands (White et al. 1998; Kaksonen et al. 2003; Labrenz & Banfield 2004). Sharma et al. (2000) studied a highly cadmium resistant Klebsiella planticola strain that is able to induce the precipitation of cadmium sulfide. This strain is a good candidate for the accelerated bioremediation of systems contaminated by high levels of cadmium. Edenborn (2004) showed that the addition of both polylactic acid (PLA) and gypsum in constructed wetlands and reactive barrier walls is required for stimulating indigenous bacteria to lower redox potential, raise pH, and enhance the bacterial sulfate reduction activity.

Phosphate precipitation. Microorganisms can produce phosphate that precipitates metals as phosphates and it is effective for a range of metals. A mixture of accumulative and chemisorptive mechanisms contribute to the overall process. Biologically produced metal precipitates as phosphates is effective for a range of metals and radionuclides (Renninger et al. 2001; Basnakova et al. 1998, See Macaskie et al. 2004 for a detailed review). A Citrobacter sp. is able to accumulate the uranyl ion (UO22+) via precipitation with phosphate liberated from organic phosphate substrates by phosphatase activity. Accumulation of uranium as HUO2PO4 by Citrobacter sp. was observed in the presence of excess inorganic phosphorus (Renninger et al. 2001). Cell-bound HUO2PO4 facilitated also Ni2+ removal by intercalative ion exchange into the polycrystalline lattice and promoted also zirconium removal (Basnakova et al. 1998). Carbonate precipitation. Bacterial calcification is another example for the precipitation of inorganic material by biofilms. The formation of calcium carbonates particles in microbial mats and other aggregates was suggested to be, at least in part, because of heterotrophic bacteria (Hammes & Verstraete 2002). Recently, interest in bacterially associated calcium carbonate precipitation has grown due to strategic bioremediation approaches involving solid phase capture of radionuclide and trace element contaminants by carbonates in groundwater systems (Curti 1999; Warren et al. 2001; Fujita et al. 2004). Fe(III) oxides are commonly thought of as the most important sorbents responsible for sequestration of many metals and radionuclides in soils and groundwater systems (Warren & Haack 2001). However, in subsurface anoxic aquifers, Fe(III) (hydr)oxides can be dissolved through microbial reactions involving reduction of Fe (III) to Fe (II). Such dissolution is accompanied by release of adsorbed or coprecipitated metals or radionuclides. Hence, if bacteria can promote precipitation of other geochemically reactive, non redox-sensitive solid phases such as carbonates, then the bacteria may provide an important bioremediation tool for solid phase sequestration of radionuclide or trace element contaminants in anoxic environments. Microbial degradation of urea was investigated as a potential geochemical catalyst for Ca carbonate precipitation and asso-

147 ciated solid phase capture of common groundwater contaminants (Sr, UO2, Cu) in laboratory batch experiments (Warren et al. 2001). Bacterial degradation of urea increased pH and promoted Ca carbonate precipitation in both bacterial control and contaminant treatments. Associated solid phase capture of Sr was highly effective, capturing 95% of the 1 mM Sr added within 24 h (Warren et al. 2001). Because bacterial ureolysis can generate high rates of calcite precipitation, the application of this approach is promising for remediation of Sr contamination in environments where calcite is stable over the long term (Fujita et al. 2004). 5.3. In-situ utilization of biofilms for bioremediation As a process for immobilizing metals, precipitation of metals as sulfides has already achieved large-scale application. Although much research is currently taking place in metal bioadsorption and the microbial reductive precipitation of metal and radionuclides contaminants, the fact is that these techniques are not yet ready for immediate use in extensive cleaning up operations. Table 3 gives some examples of bacterial strains isolated from the environment that can be used for the bioremediation of contaminated subsurface environment through microbial reductive precipitation process. However, understanding the bacterially mediated precipitation processes at a mechanistic level is still a requisite for properly applying these technologies to complex systems. Most bacteria (>90%) in subsurface or sedimentary environments are attached to particulate phases. Knowledge of the arrangement of surface-associated microorganisms is necessary in order to more fully understand the factors (e.g., electron donor and acceptor availability, pH, enzyme kinetics, diffusional limitations) that control the biogeochemistry of redox-sensitive minerals. This knowledge will allow to determine the role of bacteria in influencing or controlling the fate and transport of heavy metals and radionuclides in contaminated environments. The current understanding of sorption and accumulation of heavy metals and radionuclides by bacteria attached to mineral surfaces compromises our ability to predict microbiological immobilization of

contaminants in subsurface and sedimentary environments.

6. Summary The bioremediation of metal-contaminated soils is a multi-faceted enterprise that applies a wide variety of biological activities, from single molecules to entire macroorganisms, to the substantial problem of detoxifying hazardous land. Several techniques attempt to remove the tightly bound, slowly desorbing metals; complexation and facilitated transport by bio-surfactants, cyclodextrins and exopolysaccharides; acid leaching by the chemolithotrophic activity of Thiobacillus; reductions of arsenic and plutonium to more soluble species and reduction of mercury to volatile Hg0, methylation of selenium to volatile dimethylselenide; and extraction of cadmium, zinc, lead, selenium and other metals into living plant tissue. The reductions of selenium and chromium, in contrast, have as their goal the detoxification and immobilization of the metals in-situ, with the understanding that continuous careful management will be required to maintain reducing conditions in such environments. An important question in each of these methods is the source of the bioremedial organisms. Some of the more unusual activities, such as plutonium reduction, are known to be catalyzed by only a few microorganisms and it may be necessary to introduce specific microbes for in-situ treatment. Distributing bacterial cells through soils is itself a significant undertaking, although progress is being made in this area (Bai et al. 1997) and the survival and activity of non-native microbes in a new community is never assured. In several cases, including chromate reduction, selenate reduction, selenium volatilization and mercury volatilization, the desired activities are conveniently ubiquitous, indigenous to contaminated areas and easily stimulated by hydrogen, water, warmth, and various forms of organic matter. In-situ treatments requiring whole-cell activities thus appear most straightforward when native microorganisms can be recruited. In phytoremediation, in contrast, native plants do not often appear to be the most effective; because of the comparative ease of plant cultivation,

148 however, maintaining the desired organisms poses little difficulty. All attempts at restoration of contaminated land face the similar, internally conflicting goals of speed, cost efficiency, and generation of a clean, productive patch of earth. The earliest ‘‘brute-force’’ methods of solidification/stabiliza-

tion, excavation and landfilling, and capping were markedly constrained by the first two goals, at the expense of the third. Developing chemical technology increased the efficiency of soil washing and pump-and-treat methods, greatly improving the fate of the treated land but still requiring the addition of further hazardous

Table 3. Studies on metal–microbe interaction in soils and sediments that have impact on bioremediation of metal contamination Contaminant

Comments

References

Uranium Mine sediment Soil and sediment

Uranium

Suzuki et al. 2003

Sediment

Selenium reduction

Sediment

Chromate reduction

Subsurface environment

Uranium immobilization

Soils and sediments

Cadmium

Subsurface sediment

Uranium

Selenite polluted site Subsurface sediment Soil

Selenium

Desulfosporosinus sp. Reduced U(VI) to U(IV). This organism may be stimulated with addition of sulfate and nutrients uranium. Anaerobic Bio-oxidation of 10 mM Fe(II) and precipitation of Fe(III) oxides by Dechlorosoma sp. Resulted in rapid adsorption and removal of 55 lM uranium and 81 lM cobalt. The results of this study demonstrate the potential of this novel approach for stabilization and immobilization of heavy metals and radionuclides in the environment A selenite respiring bacterium, Bacillus selenitrieducens, produced significant levels of Se()II) and is also able to reduce selenite through Se(0) to Se ()II). The microorganism showed a great potential to immobilize selenium in anaerobic environments Consortium of marine sulfate-reducing bacteria was shown to 3+ . These indigenous SRB might reduce chromate (CrO2
4 ) into Cr have potential application in bioremediation of metal contaminated sediments. Immobilisation of hexavalent uranium [U(VI)] by biofilm of the sulfate-reducing bacterium Desulfovibrio desulfuricans G20. The hexavalent uranium was immobilized following enzymatic or chemical reduction reaction. The chemical reduction of uranium was due to the presence of dissolved sulfide species. Highly cadmium resistant Klebsiella planticola strain that is able to induce the precipitation of cadmium sulfide. This strain is a good candidate for the accelerated bioremediation of systems contaminated by high levels of cadmium Stimulation of dissimilatory metal reduction activity of an enrichment of microorganisms closely related to Pseudomonas and Desulfosporisinus sp. which promote the reductive precipitation of uranium Ralstonia metallidurans CH34 can reduce selenite to elemental red selenium, that is highly insoluble Immobilization of cobalt by sulfide precipitates

Subsurface environments

Uranium

Bio-oxidation of Fe(II)

Cobalt Mercury

Microbial reduction of soluble ionic mercury, Hg(II), to volatile Hg(0). The reduced Hg(0) can then flux out of the environment and be diluted in the atmosphere Biofilm of the sulfate reducing bacterium Desulfovibrio desulfuricans attached to hematite (a-Fe2O3) surface can accumulates both U(VI) and U(IV) and thus can limit the transport of uranium in subsurface environments

Lack et al. 2002a, b

Herbel et al. 2003

Cheung and Gu 2003

Beyenal et al. 2004

Sharma et al. 2000

Nevin et al. 2003

Roux et al. 2001 Krumholz et al. 2003 Hobman et al. 2000

Neal et al. 2004

149 substances. Research in bioremediation is now opening the virtually infinite possibilities inherent in enzymatic systems. Because biological techniques are being developed to cost as little or less than the best available conventional methods, to operate in a timely manner, employ biodegradable and (with the exception of metal-laden phytoremedial shoots) non-toxic substances, and to yield functional soils and sediments, bioremediation may soon provide the tools to restore metalcontaminated soils rapidly, inexpensively and permanently. Driven by the realization that large areas of land contaminated with metals and radionuclides cannot be economically remediated using conventional chemical approaches, significant resources have become available for the area of bioremediation of metal contamination. Supported by genomics-enabled studies ongoing in many laboratories worldwide (Lovley 2003) we can expect this research area to develop further, delivering more robust techniques for the bioremediation of metal contaminated waters, soils and sediments. Exciting developments in the use of microorganisms for the recycling of metal wastes, with the formation of novel biominerals with unique properties are also predicted for the near future.

References Abdelouas A, Lu YL, Lutze W & Nuttall HE (1998a) Reduction of U(VI) to U(IV) by indigenous bacteria in contaminated ground water. J. Contam. Hydrol. 35: 217–233 Abdelouas A, Lutze W & Nuttall EH (1998b) Chemical reactions of uranium in groundwater at a mill tailings site. J. Contam. Hydrol. 34: 343–361 Abdelouas A, Lutze W & Nuttall HE (1999) Uranium contamination in the subsurface: characterization and remediation In: Burns PC & Finch R (Eds.) Uranium: Mineralogy, Geochemistry and the Environment (pp 433–473). vol. 38 Mineralogical Society of America, Washington DC, USA Abdelouas A, Lutze W, Gong W, Nuttall EH, Strietelmeier BA & Travis BJ (2000) Biological reduction of uranium in groundwater and subsurface soil. Sci. Total Environ. 250: 21–35 Ahmann D, Roberts AL, Krumholz HF & Morel FMM (1994) Microbe grows by reducing arsenic. Nature 351: 750 Ahmann D, Krumholz LR, Hemond HF, Lovley DR & Morel FMM 1997. Microbial Mobilization of arsenic from sediments of the Aberjona Watershed. Environ. Sci. Technol. 31: 2923–2930. Anderson RT & Lovley DR (2002) Microbial redox interactions with uranium: an environmental perspective. In: Livens

FR & Keith-Roach M (Eds) Microbiology and Radioactivity (pp. 205–223). Elsevier Anderson RT, Rooney-Varga J, Gaw CV & Lovley DR (1998) Anaerobic benzene oxidation in the Fe(III)-reduction zone of petroleum-contamianted aquifers. Environ. Sci. Technol. 32: 1222–1229 Anderson RT, Vrionis HA, Ortiz-Bernad I, Resch CT, Long PE, Dayvault R, Karp K, Marutzky S, Metzler DR, Peacock A, White DC, Lowe M & Lovley DR (2003) Stimulating the in situ activity of Geobacter species to remove uranium from the groundwater of a uranium-contaminated aquifer. Appl. Environ. Microbiol. 69: 5884–5891 Andres Y, MacCordick HJ & Hubert JC (1993) Adsorption of several actinide (Th, U) and lanthanide (La, Eu, Yb) ions by Mycobacterium smegmatis. Appl. Microbiol. Biotechnol. 39: 413–417 Arp G, Reimer A & Reitner J (2001) Photosynthesis-induced biofilm calcification and calcium concentrations in Phanerozoic oceans. Science 292: 1701–1704 Bai G, Brusseau ML & Miller RM (1997) Influence of a rhamnolipid biosurfactant on the transport of bacteria through a sandy soil. Appl. Environ. Microbiol. 63: 1866– 1873 Baldi F, Minacci A, Pepi M & Scozzafava A (2001) Gel sequestration of heavy metals by Klebsiella oxytoca isolated from iron mat. FEMS Microbiol. Ecol. 36: 169–174 Barkay T & Schaefer J (2001) Metal and radionuclide bioremediation: issues, considerations and potentials. Curr. Opin. Microbiol. 4: 318–323 Barnes LJ, Janssen FJ, Sherren J, Versteegh JH, Koch RO & Scheeren PJH (1991) A new process for the microbial removal of sulphate and heavy metal from contaminated waters extracted by a geohydrological control system. Chem. Eng. Res. Des. 69: 184–186 Bartlett RJ (1991) Chromium cycling in soils and water: links, gaps and methods. Environ. Health Perspect. 92: 17–24 Basnokova G, Spencer AJ, Palsgard E & GrimeGW& Macaskie LE (1998) Identification of the nickel uranyl phosphate deposits on Citrobacter cells by electron microscopy with electron probe X ray microanalysis and by proton induced X ray emission analysis. Environ. Sci. Technol. 32: 760–765 Bender J & Phillips P (2004) Microbial mats for multiple applications in aquaculture and bioremediation. Biores. Technol. 94: 229–238 Bender J., Lee RF & Phillips P (1995) A review of the uptake and transformation of metals metalloids by microbial mats and their use in bioremediation. J. Ind. Microbial. 14: 114–118 Bender J, Lee RF, Sheppard M, Bulski K, Phillips P, Yeboah Y & Chee Wah R (2004) A waste effluent treatment system based on microbial mats for black sea bass Centrapristis strata recycled-water aquaculture. Aqua. Eng. 31: 73–82 Berne C (2004) Etude des Mecanismes de degradation du phosphate de tributyle par des bacteries. PhD. Universite de la Mediterranee, Aix-Marseille France Beveridge TJ, Makin SA, Kadurugamuwa JL & Li Z (1997a) Interactions between biofilms and the environment. FEMS Microbiol. Rev. 20: 291–303 Beveridge TJ, Hughes MN, Lee H, Leung KT, Poole RK, Savvaidis I, Silver S & Trevors JT (1997b) Metal–microbe interactions: contemporary approaches. Adv. Microbial. Physiol 38: 177–243 Beyenal H, Sani RK, Peyton BM, Dohnalkova AC, Amonette JE & Lewandowski Z (2004) Uranium immobilization by sulfatereducing biofilms. Environ. Sci. Technol. 38: 2067–2074

150 Beynal H & Lewandowski Z (2004) Dynamics of lead immobilization in sulfate reducing biofilms. Wat. Res. 38: 2726–2736 Blake RC, Shute EA & Howard GT (1994) Solubilization of minerals by bacteria: electrophoretic mobility of Thiobacillus ferrooxidans in the presence of iron. Microbiol. 60: 3349–3357 Bodek ILyman WJReehl WFRosenblatt DH. (1988) Environmental Inorganic Chemistry: Properties, Processes and Estimation Methods. Pergamon Press, Inc, Elmsford, NY Bolton H, Xun L & Girvin DC (2000) Biodegradation of synthetic chelating agents In: Lovley D (Eds.) Environmental Microbe-Metal Interactions (pp 363–383). ASM Press, Washington, D.C. Bonthrone KM, Basnakova G, Lin F & Macaskie LE (1996) Bioaccumulation of nickel byintercalation into polycrystalline hydrogen uranyl phosphate deposited via an enzymatic mechanism. Nat. Biotechnol. 14: 635–638 Bontidean I, Csoregi E, Corbisier P, Lloyd JR & Brown NL (2002) Bacterial metal- responsive elements and their use in biosensors for monitoring of heavy metals In: Sankar B (Eds.) The Handbook of Heavy Metals in the Environment (pp 647–680). Marcel Dekker Inc, New York Bontidean I, Lloyd JR, Hobman JL, Wilson JR, Csoregi E, Mattiasson B & Brown NL (2000) Bacterial metal-resistance proteins and their use in biosensors for the detection of bioavailable heavy metals. J. Inorg. Biochem. 79: 225–229 Bopp L & Ehrlich H (1988) Chromate resistance and reduction in Pseudomonas fluorescens strain LB300. Arch. Microbiol. 150: 426–431 Borst-Pauwels GWFH (1981) Ion transport in yeast. Biochim. Biophys. Acta 650: 88–127 Boseker K (1997) Bioleaching: metal solubilization by microorganisms. FEMS Microbiol. Rev. 20: 591–604 Boswell CD, Dick RE, Eccles H & Macaskie LE (2001) Phosphate uptake and release by Acinetobacter johnsonii in continuous culture and coupling of phosphate release to heavy metal accumulation. J. Ind. Microbiol. Biotechnol. 26: 333–340 Boyle V (1993) Soils washing In: Pratt M (Eds.) Remedial Processes for Contaminated Land (pp 33–53). Institution of Chemical Engineers, Warwicksjhire Brierley CL (1982) Microbiological Mining. Sci. Am. 247: 44–53 Brierley CL (1990) Biomediation of metal-contaminated surface and ground waters. Geomicrobiology 8: 201–223 Brown DA, Beveridge TJ, Keevil CW & Sherriff BL (1998a) Evaluation of microscopic techniques to observe iron precipitation in a natural microbial biofilm. FEMS Microbiol. Ecol. 26: 297–310 Brown DA, Sawicki JA & Scheriff BL (1998b) Alteration of microbially precipitated iron oxides and hydroxides. Am. Mineral. 83: 1419–1425 Brown DA, Scheriff BL, Sawicki JA & Sparling R (1999) Precipitation of iron minerals by a natural microbial consortium. Geochim. Cosmochim. Acta 63: 2163–2169 Bruins MR, Kapil S & Oehme FW (2000) Microbial resistance to metals in the environment. Ecotoxicol. Environ. Saf. 45: 198–207 Brusseau ML, Wang X & Hu Q (1994) Enhanced transport of low-polarity organic compounds through soil by cyclodextrin. Environ. Sci. Technol. 31: 952–956 Brusseau ML, Wang X & Wang WZ (1997) Simultaneous elution of heavy metals and organic compounds from soil by cyclodextrin. Environ. Sci. Technol. 31: 1087–1092

Caccavo F Jr, Lonergan DJ, Lovley DR, Davis M, Stolz JF & McInerney MJ (1994) Geobacter sulfurreducens sp. nov., a hydrogen and acetate-oxidizing dissimilatory metal reducing microorganism. Appl. Environ. Microbiol. 60: 3752–3759 Carpentier W, Sandra K, De Smet I, Brige A, De Smet L & Van Beeumen J (2003) Microbial reduction and precipitation of vanadium by Shewanella oneidensis. Appl. Environ. Microbiol. 69: 3636–3639 Cervantes C (1991) Bacterial interactions with chromate. Antonie van Leeuwenhoek 59: 229–233 Chaney RL (1997) Phytoremediation of soil metals. Cur. Opin. Biotechnol. 8: 370–384 Chen JM & Hao OJ (1998) Microbial chromium(VI) reduction. Crit. Rev. Environ. Sci. Technol. 28: 219–251 Chen SY & Lin JG (2000) Influence of solid content on bioleaching of heavy metals from contaminated sediment by Thiobacillus spp. J. Chem. Technol. Biotechnol. 75: 649–656 Chen SY & Lin JG (2001a) Bioleaching of heavy metals from sediment: significance of pH. Chemosphere 443: 1093–1102 Chen SY & Lin JG (2001b) Effect of substrate concentration on bioleaching of metal- contaminated sediment. J. Hazard. Mater. 82: 77–89 Chen SY, Lin J-G & Lee CY (2003) Effects of ferric ion on bioleaching of heavy metals from contaminated sediment. Wat. Sci. Technol. 48: 151–158 Chen S-Y & Lin J-G (2004) Bioleaching of heavy metals rom contaminated sediment by indigenous sulfur-oxidizing bacteria in an air-lift bioreactor: effects of sulfur concentration. Wat. Res. 38: 3205–3214 Chen JH, Lion LW, Ghiorse WC & Shuler ML (1995) Mobilization of adsorbed cadmium and lead in aquifer material by bacterial extracellular polymers. Wat. Res. 29: 421–430 Cheng CN & Focht D (1979) Production of arsine and methylarsines in soil and in culture. Appl. Environ. Microbiol. 38: 494–498 Cheung KH & Gu JD (2003) Reduction of chromate (CrO42)) by an enrichment consortium and an isolate of marine sulfate-reducing bacteria. Chemosphere 52: 1523–1529 Cifuentes FR, Lindemann WC & Barton LL (1996) Chromium sorption and reduction in soil with implications to bioremediation. Soil Sci. 161: 233–24 Combs GF Jr, Garbisu C, Lee BC, Yee A, Carlson DE, Smith NR, Magyarosy AC, Leighton T & Buchanan BB (1996) Bioavailability of selenium accumulated by selenite-reducing bacteria. Biol. Trace Elem. Res. 52: 209–225 Cooke T & Bruland K (1987) Aquatic chemistry of selenium: evidence of biomethylation. Environ. Sci. Technol. 21: 1214–1219 Couillard D & Mercier G (1991) Optimum residence time (in CSTR and airlift reactors) for bacterial leaching of metals from anaerobic sewage sludge. Wat. Res. 25: 211–218 Couillard D & Mercier G (1993) Removal of metals and fate of N and P in the bacterial leaching of aerobically digested sewage sludge. Wat. Res. 27: 1227–1235 Cox JS, Scott Smith D, Warren LA & Ferris FG (1999) Characterizing heterogeneous bacterial surface functional groups using discrete affinity spectra for proton binding. Environ. Sci. Technol. 33: 4514–4521 Cummings DE, Snoeyenbos-West OL, Newby DT, Niggemyer AM, Lovley DR, Achenbach LA & Rosenzweig RF (2003) Diversity of Geobacteraceae species inhabiting metal-polluted freshwater lake sediments ascertained by 16S rDNA analyses. Microb. Ecol. 46: 257–269

151 Cunningham SD, Berti WR & Huang JW (1995) Phytoremediation of contaminated soils. Trends Biotechnol. 13: 393–397 Curti E (1999) Coprecipitation of radionuclides with calcite: estimation of partition coefficients based on a review of laboratory investigations and geochemical data. Appl. Geochem. 14: 433–445 Daly MJ (2000) Engineering radiation-resistant bacteria for environmental biotechnology. Curr. Opin. Biotechnol. 11: 280–285 Davison W, Fones GR & Grime GW (1997) Dissolved metals in surface sediment and a microbial mat at 100lm resolution. Nature 387: 885–888 DeMoll-Decker HM & Macy JM (1993) The periplasmic nitrite reductase of Thanera selenatis may catalyze the reduction of selenite to elemental selenium. Arch. Microbiol. 160: 241–247 deRome L & Gadd GM (1991) Use of pelleted and immobilized yeast and fungal biomass for heavy metal and radionuclide recovery. J. Ind. Microbiol 7: 97–104 Diels L, Dong Q, van der Lelie D, Baeyens W & Mergeay M (1995) The czc operon of Alcaligenes eutrophus CH34: from resistance mechanism to the removal of heavy metals. J. Ind. Microbiol. 14: 142–153 Diels L, Spaans H, Van Roy S, Hooyberghs L, Ryngaert A, Wouters H, Walter E, Winters J, Macaskie L, Finlay J, Pernfuss B, Woebking H, Pu¨mpel T & Tsezos M (2003) Heavy metals removal by sand filters inoculated with metal sorbing and precipitating bacteria. Hydrometallurgy 71: 235–241 Douglas S & Beveridge TJ (1998) Mineral formation by bacteria in natural microbial communities. FEMS Microbiol. Ecol. 26: 79–88 Drzyzga O, El Mamouni R, Agathos SN & Gottschal J (2002) Dehalogenation of chlorinated ethenes and immobilization of nickel in anaerobic sediment columns under sulfidogenic conditions. Environ. Sci. Technol. 36: 2630–2635 Eccles H, Garnham GW, Lowe CR & Bruce NC (1997) Biosensors for detecting metal ions capable of being reduced by reductase enzymes. Biotechnol. Adv. 15: 189–189 Edenborn HM (2004) Use of poly(lactic acid) amendments to promote the bacterial fixation of metals in zinc smelter tailings. Biores. Technol. 92: 111–119 Ellwood DC, Hill MJ & Watson JHP (1992) Pollution control using microorganisms and metal separation In: Fry JC, Gadd GM, Herbert RA, Jones CW & Watson-Craik A (Eds.) Microbial Control of Polution (pp 89–112). Vol. 48Cambridge University Press, Cambridge, U.K Fendorf S, Wielinga BW & Hansel CM (2000) Chromium transformations in natural environments: the role of biological and abiological processes in chromium(VI) reduction. Intl. Geol. Rev. 42: 691–701 Ferris FG, Schultze , Witten T, Fyfe WS & Beveridge TJ (1989) Metal interactions with microbial biofilms in acidic and neutral pH environments. Appl. Environ. Microbiol. 55: 1249–1257 Finneran KT, Anderson RT, Nevin KP & Lovley DR (2002) Potential for bioremediation of uranium-contaminated aquifers with microbial U(VI) reduction. Soil Sed. Comtam. 11: 339–357 Flemming H-C, Leis A (2002) Sorption properties of biofilms. In: Flemming H-C (Ed) Biofilms. In: Bitton G. (Ed) Encyclopedia of Environmental Microbiology, vol. 5, (pp. 2958–2967) Fo¨rstner U (2004) Sediment dynamics and pollutant mobility in rivers: An interdisciplinary approach. Lakes Reservoirs Res. Manag. 9: 25–40

Fortin D (2004) Geochemistry. What biogenic minerals tell us. Science 303: 618–1619 Francis AJ (1994) Microbial transformations of radioactive wastes and environmental restoration through bioremediation. J. Alloys Comp 213/214: 226–231 Francis AJ & Dodge CJ (1998) Remediation of soils and wastes contaminated with uranium and toxic metals. Environ. Sci. Technol. 32: 3993–3998 Frankenberger WT Jr. & Karlson U (1994a) Microbial volatilization of selenium from soils and sediments In: Frankenberger WT Jr. & Benson S (Eds.) Selenium in the Environment (pp 369–387). Marcel Dekler, Inc., New York Frankenberger WT Jr. & Karlson U (1994b) Soil management factors affecting volatilization of selenium from dewatered sediments. Geomicrobiol. J. 12: 265–278 Frankenberger WT Jr & Karlson U (1995) Volitilization of selenium from a dewatered seleniferous sediment: a field study. J. Indust. Microbiol. 14: 226–232 Fredrickson JK, Zachara JM, Kennedy DW, Lin C, Duff MC, Hunter DB & Dohnalkova A (2002) Influence of Mn oxides on the reduction of uranium (VI) by the metal-reducing bacterium Shewanella putrefaciens. Geochem. Vosmochim. Acta 66: 3247–3262 Fredrickson JK, et al. (2000) Reduction of Fe(III), Cr(VI), U(VI), and Tc(VII) by Deinococcus radiodurans. Appl. Environ. Microbiol. 66: 2006–2011 Fujita Y, Redden GD, Ingram JC, Cortez MM, Ferris FG & Smith RW (2004) Strontium incorporation into calcite generated by bacterial ureolysis. Geochim. Cosmochim. Acta 68: 3261–3270 Gadd GM (2004) Microbial influence on metal mobility and application for bioremediation. Geoderma 122: 109–119 Gadd GM (2000) Microbial interactions with tributyltin compounds; detoxification, accumulation, and environmental fate. Sci. Total. Environ. 258: 119–127 Gadd GM & White C (1989) Heavy metal and radionuclide accumulation and toxicity in fungi and yeasts In: Poole RK & Gadd GM (Eds.) In Metal–Microbe Interactions. Society for General Microbiology (pp 19–38). IRL Press, Oxford, England Geesey GG, Lang L, Jolley JG, Hankins MR, Iwaoka T & Griffiths PR (1989) Binding of metal ion by extracellular polymers of biofilm bacteria. Water Sci. Technol. 20: 161–165 Geiselhart L, Osgood M & Holmes DS (1991) Ann. NY Acad. Sci. 646: 53–60 Glazer BT, Cary SC, Hohmann L & Luther GW III (2002) In situ sulfur speciation using Au/Hg microelectrode as an aid to microbial characterisation of an intertidal salt marsh microbial mat In: Taillefert M & Rozan T (Eds.) Environmental Electrochemistry: Analyses of Trace Element Biogeochemistry (pp 283–305). American Chemical Society Symposium Series; American Chemical Society, Washington D.C Gorby YA & Lovley DR (1992) Enzymatic uranium precipitation. Environ. Sci. Technol. 26: 205–207 Gorby YACF and Bolton H (1998) Microbial reduction of cobalt(III)EDTA in the presence and absence of manganese(IV) oxide. Environ. Sci. Technol. 32: 244–250 Greiner R, Konierny & Jany KD (1993) Purification and characterization of two phytases from Eschericha Coli. Arch. Biochem. Biophys. 303: 107–113 Haack EA & Warren LA (2003) Biofilm hydrous manganese oxyhydroxides and metal dynamics in acid rock drainage. Environ. Sci. Technol. 37: 4138–4147

152 Hammes F & Verstraete W (2002) Key roles of pH and calcium metabolism in microbial carbonate precipitation. Re/Views Environ. Sci. Bio/Technol. 1: 3–7 Herbel MJ, Switzer Blum J, Borglin SE & Oremland RS (2003) Reduction of elemental selenium to selenide: experiments with anoxic sediments and bacteria that respire Se- oxyanions. Geomicrobiol. J. 20: 587–602 Herman DC, Artiola JF & Miller RM (1995) Removal of cadmium, lead and zinc from soil by a rhamnolipid biosurfactant. Environ. Sci. Technol. 29: 2280–2285 Higham DP, Sadler PJ & Scawen MD (1984) Cadmiumresistant Pseudomonas putida synthesizes novel cadmium binding proteins. Science 225: 1043–1046 Hinck ML, Ferguson J & Puhakka J (1997) Resistance of EDTA and DTPA to aerobic biodegraduation. Water Sci. Technol. 35: 25–31 Hobman JL, Wilson JR & Brown NL (2000) Microbial mercury reduction In: Lovley DR (Eds.) Environmental Microbe–Metal Interactions (pp 177–197). ASM Press, Washington, D.C Hockin SL & Gadd GM (2003) Linked redox precipitation of sulfur and selenium under anaerobic conditions by sulfatereducing bacterial biofilms. Appl. Environ. Microbiol. 69: 7063–7072 Holmes DE, Finneran KT & Lovley DR (2002) Enrichment of Geobacteraceae associated with stimulation of dissimilatory metal reduction in uranium-contaminated aquifer sediments. Appl. Environ. Microbiol. 68: 2300–2306 Jackson M, West M & Leppard GG (1999) Accumulation of heavy metals by individually analysed bacterial cells and associated nonliving material in polluted lake sediments. Environ. Sci. Technol. 33: 3795–3801 Ji G & Silver S (1995) Bacterial resistance mechanisms for heavy metals of environmental Citrobacter concern. J. Indust. Microbiol. 14: 61–75 John SG (2001) Siderophore mediated plutonium accumulation by Microbacterium flavescens (JG-9). Environ. Sci. Technol. 35: 2942–2948 Kaksonen AH, Riekkola-Vanhanen ML & Puhakka JA (2003) Optimization of metal sulfide precipitation in fluidised-bed treatment of acidic wastewater. Water Res. 37: 255–266 Kearney T, Eccles H, Graves D & Gonzalez A (1996) Presented at the Proceedings of the 18th Annual Conference of the National Low-Level Waste Management Program, May 20– 22, Salt Lake City, Utah, USA Kjaergaard K, Sorenson JK, Schembri MA & Klemm P (2000) Sequestration of zinc oxide by fimbrial designer chelators. Appl. Environ. Microbiol. 66: 10–14 Klaus T, Joeger R, Olsson E & Granqvist CG (1999) Silverbased crystalline nanoparticles, microbially fabricated. Proc. Natl. Acad. Sci. USA 96: 13611–13614 Konhauser KO, Fyfe WS, Ferris FG & Beveridge TJ (1993) Metal sorption and mineral precipitation by bacteria in two Amazonian river systems: Rio Solimoes and Rio Negro, Brazil. Geology 21: 1103–1106 Konhauser KO, Schultze-Lam S, Ferris FG, Fyfe WS, Longstaffe FJ & Beveridge TJ (1994) Mineral precipitation by epilithic biofims in the Speed River, Ontario, Canada. Appl. Environ. Microbiol. 60: 549–553 Kostka JE (2002) Growth of iron(III)-reducing bacteria in clay minerals as the sole electron acceptor and comparison of growth yields on a variety of oxidized iron forms. Appl. Environ. Microbiol. 68: 6256–6262

Krishnamurthy S & Wilkens MM (1994) Environmental chemistry of chromium. Northeast. Geol. 16: 14–17 Krumbein WE (1979) Calcification by bacteria and algae. In: Trudinger PA & Swaine DJ (Eds) Biogeochemical Cycling of Mineral- Forming Elements. Elsevier Science, New York Krumholz LR, Elias DA & Suflita JM (2003) Immobilization of cobalt by sulfate-reducing bacteria in subsurface sediments. Geomicrobiol. J. 20: 61–72 Kuhn A & Sigg L (1993) Arsenic cycling in eutrophic Lake Greifen, Switzerland: influence of seasonal redox processes. Limnol. Oceanogr. 38: 1052–1055 Labrenz M & Banfield JF (2004) Sulfate reducing bacteriadominated biofilms that precipitate ZnS in a subsurface circumneutral-pH mine drainage system. Microb. Ecol. 47: 205–217 Labrenz M, Druschel GK, Thomsen-Ebert T, Gilbert B, Welch SA, Kemner KM, Logan GA, Summons RE, De Stasio G, Bond PL, Lai B, Kelly SD & Banfield JF (2000) Formation of sphalerite (ZnS) deposits in natural biofilms of sulfatereducing bacteria. Science 290: 1744–1747 Lack JG, Chaudhuri SK, Kely SD, Kemmer KM, O’Connor SM & Coates JD (2002) Immobilization of radionuclides and heavy metals through anaerobic bio-oxidation of Fe(II). Appl. Environ. Microbiol. 68: 2704–2710 Lack JG, et al. (2002) Immobilization of radionuclides and heavy metals through anaerobic bio-oxidation of Fe(III) by Dechlorosoma suillum. Microbiol. Ecol. 43: 424–431 Langley S & Beveridge TJ (1999) Effect of O-side-chainlipopolysaccharide chemistry on metal binding. Appl. Environ. Microbiol. 65: 489–498 Lappin-Scott HM, Jass J & Costerton JW (1993) Microbial biofilm formation and characterization In: Lappin-Scott HM & Costerton JW (Eds.) Microbial Films: Formation and Control (pp 1–73). SAB Press, London, UK Laverman AM, Blum JS, Schaefer JK, Phillips EJP, Lovley DR & Oremland RS (1995) Growth of strain SES-3 with arsenate and other diverse electron acceptors. Appl. Environ. Microbiol. 61: 3556–3561 Lawson S & Macy J (1995) Bioremediation of selenite in oil refinery wastewater. Appl. Microbiol. Biotechnol. 43: 762–765 Liehr SK, Chen HJ & Lin SH (1994) Metals removal by algal biofilms. Water Sci. Technol. 30: 59–60 Lloyd JR (2003) Microbial reduction of metals and radionuclides. FEMS Microbiol. Rev. 27: 411–425 Lloyd JR & Macaskie LE (1996) A novel phosphorImager based technique for monitoring the microbial reduction of technetium. Appl. Environ. Microbiol. 62: 578–582 Lloyd JR & Macaskie LE (2000) Bioremediation of radioactive metals In: Lovley DR (Eds.) Environmental Microbe–Metal Interactions (pp 277–327). ASM Press, Washington, D.C Lloyd JR & Lovley DR (2001) Microbial detoxification of heavy metals and radionuclides. Curr. Opin. Biotechnol. 12: 248–253 Lloyd JR & Macaskie LE 2002. The biochemical basis of radionuclide-microbe Interaction. In: FR Livens and M Keith-Roach (ed.), Microbiology and Radioactivity. Elsevier Lloyd JR, Sole VA, Van Praagh CV & Lovley DR (2000) Direct and Fe(II)-mediated reduction of technetium by Fe(III)-reducing bacteria. Appl. Environ. Microbiol. 66: 3743–3749 Lloyd JR, Mabbett AN, Williams DR & Macaskie LE (2001) Metal reduction by sulfate-reducing bacteria: physiological diversity and metal specificity. Hydrometallurgy 59: 327–337

153 Loser C, Seidel H, Hoffman P & Zehnsdorf A (2001) Remediation of heavy metal-contaminated sediments by solid-bed bioleaching. Environ. Geol. 40: 643–650 Losi ME & Frankenberger WT (1997) Bioremediation of selenium in soil and water. Soil Sci. 162: 692–702 Losi ME, Amrhein C & Frankenberger WT Jr (1994a) Factors affecting chemical and biological reduction of hexavalent chromium in soil. Environ. Toxicol. Chem. 13: 1727–1735 Losi ME, Amrhein C & Frankenberger WT Jr. (1994b) Bioremediation of chromate-contaminated groundwater by reduction and precipitation in surface soils. J. Environ. Qual. 23: 1141–1150 Lovley DR (1993) Dissimilatory metal reduction. Ann. Rev. Microbiol. 47: 263–290 Lovley D (2000) Environmental Microbe–Metal Interactions. ASM Press, Washington, D.C. 395 Lovley DR (2001a) Reduction of iron and humics in subsurface environments In: Frederickson JK & Fletcher M (Eds.) Subsurface Microbiology and Biogeochemistry (pp 193–217). Wiley-Liss, Inc., New York Lovley DR (2001b) Biomediation – anerobes to the rescue. Science 293: 1444–1446 Lovley D (2002) Dissimilatory metal reduction: from early life to biomediation. ASM News 68: 231–237 Lovley DR (2003) Cleaning up with genomics: Applying molecular biology to bioremediation. Nat. Rev. Microbiol. 1: 36–44 Lovley D & Phillips EJ (1992a) Reduction of uranium by Desulfovibrio desulfuricans. Appl. Environ. Microbiol 58: 850–856 Lovley DR & Phillips EJP (1992b) Bioremediation of uranium contamination with enzymatic uranium reduction. Environ. Sci. Technol. 26: 2228–2234 Lovley DR & Blunt-Harris EL (1999) Role of humics-bound iron as an electron transfer agent in dissimilatory Fe(III) reduction. Appl. Environ. Microbiol. 65: 4252–4254 Lovley DR & Coates JD (2000) Novel forms of anaerobic respiration of environmental relevance. Curr. Opin. Microbiol. 3: 252–256 Lovley DR, Phillips EJP, Gorby YA & Landa E (1991) Microbial reduction of uranium. Nature 350: 413–416 Lovley DR, Roden EE, Phillips EJP & Woodward JC (1993a) Enzymatic iron and uranium reduction by sulfate reducing bacteria. Marine Geol. 113: 41–53 Lovley DR, Widman PK, Woodward JC & Phillips EJP (1993b) Reduction of uranium by cytochrome c3 of Desulfovbrio vulgaris. Appl. Environ. Microbiol. 59: 3572–3576 Lovley DR & Phillips EJP (1994) Reduction of chromate by Desulfovibrio vulgaris and its cytochrome. Appl. Environ. Microbiol. 60: 726–728 Macaskie LE (1991) The application of biotechnology to the treatment of wastes produce from nuclear fuel cycle: biodegradation and bioaccumulation as a means of treating radionuclide-containing streams. Crit. Rev. Biotechnol. 11: 41–112 Macaskie LE, Empson RM, Cheetham AK, Grey CP & Skarnulis AJ (1992) Uranium bioaccumulation by a Citrobacter sp. as a result of enzymically-mediated growth of polycrystalline HUO2PO4. Science 257: 782–784 Macaskie LE, Jeong BC & Tolley MR (1994) Enzymicallyaccelerated biomineralization of heavy metals: application to the removal of americium and plutonium from aqueous flows. FEMS Microbiol. Rev. 14: 351–379 Macaskie LE, Lloyd JR, Thomas RAP & Tolley MR (1996) The use of microorganisms for the remediation of solutions

contaminated with actinide elements, other radionuclides, and organic contaminants generated by nuclear fuel cycle activities.. Nucl. Energy 35: 257–271 Macaskie LE, Yong P & Paterson-Beedle M (2004) Bacterial precipitation of metal phosphates In: Valsami-jones E (Eds.) Phosphorus in Environmental Technology; Principles and Applications (pp 547–579). IWA publishing, London, UK Macaskie LE (1991) The application biotechnology to the treatment of wastes produced from the nuclear fuel cycle: biodegradation and bioaccumulation as a means of treating radionuclide-containing streams. Crit. Rev. Biotechnol. 11: 41–112 Macy J (1994) Biochemistry of selenium metabolism by Thauera selenatis gen.nov. sp. Nov and use of the organism for bioremediation of selenium oxyanion in San Joaquin Valley drainage water In: Frankenberger WT Jr. & Benson S (Eds.) Selenium in the Environment (pp 421–444). Marcel Dekker, Inc., New York, NY Macy J, Lawson S & DeMoll-Decker H (1993) Bioremediation of selenium oxyanions in San Joaquin drainage water using Thauera selentis in a biological reactor system. Appl. Microbiol. 40: 588–594 Macy J, Nunan K, Hagen K, Dixon D, Harbour P, Cahill M & Sly L (1996) Chryioenes arsenatis gen. nov., sp. Nov., a new arsenate-respiring bacterium isolated from gold mine wastewater. Int. J. Syst. Bacteriol. 46: 1153–1157 McCullough J, Hazen TC, Benson SM, Metting FB & Palmisano AC (1999) Bioremediation of Metals and Radionuclides. What Is It and How it Works. Lawrence Berkeley National Laboratory, Berkeley, Cailfornia McHale AP & McHale S (1994) Microbial biosorption of metals: potential in the treatment of metal pollution. Biotechnol. Adv. 12: 647–652 Mercier G, Chartier M & Couillard D (1996) Strategies to maximize the microbial leaching of lead from metal-contaminated aquatic sediments. Wat. Res. 30: 2452–2464 Miller RM (1995) Biosurfactant-facilitated remediation of metal-contaminated soils. Environ. Health Perspect. 103(Suppl. 1), 59–62 Morel FMM & Hering JG (1993) Principles and Applications of Aquatic Chemistry. John Wiley & Sons, Inc, New York, NY Neal AL, Amonette JE, Peyton BM & Geesey GG (2004) Uranium complexes formed at hematite surfaces colonized by sulfate-reducing bacteria. Environ. Sci. Technol. 38: 3019–3027 Nelson YM, Lion LW, Shuler ML & Ghiorse WC (2002) Effect of oxide formation mechanisms on lead adsorption by biogenic manganese hydroxides, iron hydroxides, and their mixtures. Environ. Sci. Technol. 36: 421–425 Neu MP (2000) Siderophore-mediated chemistry and microbial uptake of plutonium. Los Alamos Sci. 26: 416–417 Neu MP, Matonic JH, Ruggiero CE, Scott BL & Angew (2000b) The first structural characterization of plutonium(IV) complexed by a siderophore: Single crystal structure of Pu Desferrioxamine E. Angewandte Chemie Intl. Edit. 39: 1442– 1444 Nevin KP, Finneran KT & Lovley DR (2003) Microorganisms associated with Uranium bioremediation in a high-salinity subsurface sediment. Appl. Environ. Microbiol. 69: 3672– 3675 Newman DK & Banfield JF (2002) Geomicrobiology: how molecular-scale interactions underpin biogeochemical systems. Science 296: 1071–1077

154 Newman D, Kennedy E, Coates J, Ahmann D, Ellis D, Lovley D & Morel F. (1997) Dissimilatory As(V) and Se(VI) reduction in Desulfotomaculum auripigmentum, sp. Nov. Arch. Microbiol. 168: 380–388 North NN, Dollhopf SL, Petrie L, Istok JD, Balkwill DL & Kostka JE (2004) Change in bacterial community structure during in situ biostimulation of subsurface sediment cocontaminated with uranium and nitrate. Appl. Environ. Microbiol. 70: 4911–4920 Okorov LA, Lichko LP, Kodomtseva VM, Kholodenko VP, Titovsky VT & Kulaev IS (1977) Energy-dependent transport of manganese into yeast cells and distribution of accumulated ions. Eur. J. Biochem. 75: 373–377 Oremland R, Blum JS, Culbertson C, Visscher P, Miller L, Dowdle P & Strohmaier F (1994) Isolation, growth and metabolism of an obligately anaerobic, selenate-respiring bacterium, strain SES-3. Appl. Environ. Microbiol. 60: 3011– 3019 Oremland R, Steinberg N, Maest A, Miller L & Hollibaugh J (1990) Measurement of removal of in situ rates of selenate removal by dissimilatory bacterial reduction in sediments. Environ. Sci. Technol. 24: 1157–1164 Oremland RS, Hollibaugh JT, Maest AS, Presser TS, Miller LG & Culbertson CW (1989) Selenate reduction to elemental selenium by anaerobic bacteria in sediments and culture: biogeochemical significance of a novel, sulphate-independent respiration. Appl. Environ. Microbiol. 55: 2333–2343 Ortiz-Bernad I, Anderson RT, Vrionis HA & Lovley DR (2004) Vanadium respiration byGeobacter metallireducens: a novel strategy for the in situ removal of vanadium from groundwater. Appl. Environ. Microbiol. 70: 3091–3095 Paterson-Beedle M & Macaskie LE (2003) Presented at the IBS 2003 Biohydrometallurgy: A sustainablee Technology in Evolution, Athens, Greece, September 14–17th Pattanapipitpaisal P, Mabbett AN, Finlay JA, Beswick AJ, Paterson-Beedle M, Wright EAJ, Tolley MR, Badar U, Ahmed N, Hobman JL, Brown NL & Macaskie LE (2002) Reduction of Cr(VI) and bioaccumulation of chromium by Gram positive and Gram negative microorganisms not previously exposed to Cr-stress. Environ. Technol. 23: 731–745 Payne RB, et al. (2002) Uranium reduction by Desulfiovbrio desulfuricans strain G20 and a cytochrome c3 mutant. Appl. Environ. Microbiol. 68: 3129–3132 Pazirandeh M, Wells BM & Ryan RL (1998) Development of bacterium-based heavy metal biosorbants : enhanced uptake of cadmium and mercury by Escherichia coli expressing a metal-binding motif. Appl. Environ. Microbiol. 64: 4068– 4072 Phillips EJP, Landa ER & Lovley DR (1995) Remediation of uranium contaminated soils withbicarbonate extraction and microbial U(VI) reduction. J. Indust. Microbiol. 14: 203–20 Postgate JR (1979) The Sulphate Reducing Bacteria. Cambridge University Press, Cambridge, U.K Pozas-Tormo R, Moreno-Real L, Martinez L & BruqueGomez S (1987) Intercalation of lanthanides into H3OUO2PO4Æ3H2O and C4H9NH3UO2PO4Æ3H2O. Inorg. Chem. 26: 1442–1445 Pulford ID & Watson C (2003) Phytoremediation of heavy metyal-contaminated land by trees: a review. Environ. Int. 29: 539–560 Raskin I, Smith RD & Salt DE (1997) Phytoremediation of metals: Using plants to remove pollutants from the environment. Curr. Opin. Biotechnol. 8: 221–226

Reeves RD & Baker AJM (2001) Metal-accumulating plants In: Raskin I & Ensley BD (Eds.) Phytoremediation of Toxic Metals: Using Plants to Clean-up the Environment (pp 193– 221). John Wiley and Sons, New York Remoudaki E, Hatzikioseyian A, Kousi P & Tsezos M (2003) The mechanism of metals precipitation by biologically generated alkalinity in biofilm reactors. Water Res. 37: 3843–3854 Renninger N, McMahon KD, Knopp R, Nitsche H, Clark DS & Keasling JD (2001) Uranyl precipitation by biomass from an enhanced biological phosphorus removal reactor. Biodegradation 12: 401–410 Rodriguez H & Fraga R (1999) Phosphate solubilizing bacteria and their role in plant growth promotion. Biotechnol. Adv. 17: 319–339 Roig MG, Kennedy JF & Macaskie LE (1995) Biological Rehabilitation of Metal Bearing Wastewaters. Final report EC contract EV5V-CT93–0251. Programme: Environment Ro¨ling WFM, van Breukelen BM, Braster M, Lin B & van Verseveld HW (2001) Relationships between microbial community structure and hydrochemistry in a landfill leachatepolluted aquifer. Appl. Environ. Microbiol. 67: 4619–4629 Roux M, Sarret G, Pignot-Paintrand I, Fontecave M & Coves J (2001) Mobilization of selinite by Ralstonia metallidurans CH34. Appl. Environ. Microbiol. 67: 769–773 Rugh CL, Wilde HD, Stack NM, Thompson DM, Summers AO & Meagher RB (1996) Mercuric ion reduction and resistance in transgenic Arabidopsis thaliana plants expressing a modified bacteria merA gene. Proc. Natl. Acad. Sci. 93: 3182–3187 Rusin P, Quintana L, Brainard J, Strietelmeier B, Tait C, Ekberg S, Palmer P, Newton T & Clark D (1994) Solubilization of plutonium hydrous oxide by non-reducing bacteria. Environ. Sci. Technol. 28: 1686–1690 Salt DE, Blaylock M, Kumar NPBA, Dushenkov V, Ensley BD, Chet I & Raskin I (1995) Phytoremediation: a novel strategy for the removal of toxic metals from the environment using plants. Biotechnol. 13: 468–474 Salt DE, Smith RD & Raskin I (1998) Phytoremediation. Ann. Rev. Plant Phys. Plant Mol. Biol. 49: 643–668 Samuelson P, Wernerus H, Svedberg M & Stahl S (2000) Staphylococcal surface display of metal-binding polyhistidyl peptides. Appl. Environ. Microbiol. 66: 1243–1248 Sand W, Gerke T, Hallmann R & Schippers A (1995) Sulfur chemistry, biofilm and the indirect attack mechanism-acritical evaluation of bacterial leaching. Appl. Microbiol. Biotechnol. 43: 961–966 Satroutdinov AD, Dedyukhina EG, Chistyakova TI, Witschel M, Minkevich IG, Eroshin VK & Egli T (2000) Degradation of metal–EDTA complexes by resting cells of the bacterial strain DSM 9103. Environ. Sci. Technol. 34: 1715– 1720 Saxena P, Krishma Raj S, Dan T, Perras M & Vettakkorumakankav N (1999) Phytoremediation of heavy metal contaminated and polluted soils In: Prasad MNV & Hagener J (Eds.) Heavy Metal Stress in Plants; from Molecules to Ecosystems (pp 305–329). Springer-Verlag, Berlin, Heidelberg Schleck DS (1990) In situ treatment: When does it apply? Superfund ’90: Proceedings of the National Conference: Nov. 26–28 (pp. 677–680), Environmental Protection Agency, Washington, D.C

155 Schultze-Lam S, Fortin D, Davis BS & Beveridge TJ (1996) Mineralization of bacterial surfaces. Chem. Geol. 132: 171–181 Schwitzguebel JP, van der Lelie D, Baker A, Glass D & Vangrosweld J (2002) Phytoremediation: european and american trends, successes, obstacles and needs. J. Soils Sediments 2: 91–99 Seidel H, Loser C, Zensdorf A, Hoffman P & Scmerold R (2004) Bioremediation process for sediments contaminated by heavy metals: feasibility study on a pilot scale. Environ. Sci. Technol. 38: 1582–1588 Sharma PK, Balkwill DL, Frenkel A & Vairavamurthy MA (2000) New Klebsiella planticola strain (Cd-1) grows anaerobically at high cadmium concentrations and precipitates cadmium sulfide. Appl. Environ. Microbiol. 66: 3083–3087 Shen H & Wang YT (1993) Characterization of enzymatic reduction of hexavalent chromium by Escherichia coli ATCC 33456. Appl. Environ. Microbiol. 59: 3771–3777 Sheppard SC & Evenden WG (1992) Bioavailability indices for uranium: effect of Concentration in eleven soils. Arch. Environ. Contam. Toxicol. 23: 117–124 Silver S (1994) Exploiting heavy metal resistance systems in bioremediation. Res. Microbiol. 145: 61–67 Smith WL & Gadd GM (2000) Reduction and precipitation of chromate by mixed culture sulfate- reducing bacterial biofilms. J. Appl. Microbiol. 88: 983–991 Smucker SJ (1996) Region 9 preliminary remediation goals; http://www.epa.gov/region9/waste/sfund/prg/index.html:U.S Snoeyenbos-West O, Nevin KP, Anderson RT & Lovley DR (2000) Enrichment of Geobacter species in response to stimulation of Fe(III) reduction in sandy aquifer sediments. Microbiol. Ecol. 39: 153–167 Steinberg N, Blum J, Hochstein L & Oremland R (1992) Nitrate is a preferred electron acceptor for growth of freshwater selenate-respiring bacteria. Appl. Environ. Microbiol. 501: 426–428 Stumm W & Morgan JJ (1996) Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters, 3. J. Wiley & Sons, New York, 1022 pp Suzuki Y, Kelly SD, Kemner KM & Banfield JF (2003) Microbial populations stimulated for hexavalent uranium reduction in uranium mine sediment. Appl. Environ. Microbiol. 69: 1337–1346 Tan H, Champion JT, Artiola JF, Brusseau ML & Miller RM (1994) Complexation of cadmium by a rhamnolipid biosurfactant. Environ. Sci. Technol. 28: 2402–1413 Tebo BM & Obraztsova AY (1998) Sulfate-reducing bacterium grows with Cr(VI), U(VI), Mn(IV), and Fe(III) as electron acceptors. FEMS Micro. Lett. 162: 193–198 Templeton AS, Trainor TP, Spormann AM & Brown Jr. GE (2003) Selenium speciation and partitioning within Burkholderia cepacia biofilms formed on – Al2O3 surfaces. Geochim. Cosmochim. Acta 67: 3547–3557 Templeton AS, Trainor TP, Spormann AM, Newville M, Sutton SR, Dohnalkova A, Gorby Y & Brown Jr GE (2003) Sorption versus biomineralization of Pb(II) within Burkholderia cepacia biofilms. Environ. Sci. Technol. 37: 300–307 Thomas RAP, Beswick AJ, Basnakova G, Moller R & Macaskie LE (2000) Growth of naturally-occurring microbial isolates in metal–citrate medium and bioremediation of metal–citrate wastes. J. Chem. Technol. Biotechnol. 75: 187–195 Thomas RAP & Macaskie LE (1996) Biodegradation of tributyl phosphate by naturally occuring microbial isolates and

coupling to the removal of uranium from aqueous solution. Environ. Sci. Technol. 30: 2371–2375 Thomas RAP & Macaskie LE (1998) The effect of growth conditions on the biodegradation of tributyl phosphate and potential for the remediation of acid mine drainage waters by a naturally-occurring mixed microbial culture. Appl. Microbiol. Biotechnol. 49: 202–209 Thomas RAP, Lawlor K, Bailey M & Macaskie LE (1998) The biodegradation of metal–EDTA complexes by an enriched microbial population. Appl. Environ. Microbiol. 64: 1319– 1322 Thompson-Eagle E, Frankenberger WT Jr. & Karlson U (1989) Volatilization of selenium by Alternaria alternata. Appl. Environ. Microbiol. 55: 1406–1413 Tichy R, Rulkens WH, Grutenhuis JTC, Nydl V, Cuypers C & Fajtl J (1998) Bioleaching of metals from soils or sediments. Water Sci. Technol. 37: 119–127 Tobin JM, White C & Gadd GM (1994) Metal accumulation by fungi: applications in environmental biotechnology. J. Ind. Microbiol. 13: 126–130 Treen-Sears ME, Volesky B & Neufeld RJ (1984) Ion exchange/ complexation of the uranyl ion by Rhizopus biosorbent. Biotechnol. Bioeng. 26: 1323–1329 Trezek GJ (1990) Heavy metal-contaminated soil remediation at high throughput; Superfund ’90 (pp. 673–676): Proceedings of the National Conference, Nov. 26–28, Environmental Protection Agency, Washington, D.C Tucker MD, Barton LL & Thomson BM (1997) Reduction and immobilization of molybdenum by Desulfovibrio desulfuricans. J. Environ. Qual. 26: 1146–1152 Turick CE, Apel WA & Carmiol NS (1996) Isolation of hexavalent chromium-reducing anaerobes from hexavalent chromium-contaminated and noncontaminated environments. Appl. Microbiol. Biotechnol. 44: 683–688 Turner JS & Robinson NJ (1995) Cyanobacterial metallothioneins: Biochemistry and molecular genetics. J. Ind. Microbiol. 14: 119–125 Tsai LJ, Yu KC, Chen SF & Kung PY (2003) Effect of temperature on removal of heavy metals from contaminated river sediments via bioleaching. Water Res. 37: 2449– 2457 U.S. Department of Energy, GJO (1999) Final site observational work plan for the UMTRA project Old Rifle site GJO-99-88-TAR U.S. Department of Energy, O. o. B. a. ER (2004) posting date. Natural and Accelerated Bioremediation Research (NABIR) program. [Online.] Utgikar VP, Harmon SM, Chaudhary N, Tabak HH, Govind R & Haines JR (2002) Inhibition of sulfate reducing bacteria by metal sulfide formation in bioremediation of acid minedrainage. Environ. Toxicol. 17: 40–48 Valls M & de Lorenzo V (2002) Exploiting the genetic and biochemical capacities of bacteria for remediation of heavy metal pollution. FEMS Microbiol. Rev. 26: 327–338 Valls M., Atrian S, de Lorenzo V & Fernandez LA (2000) Engineering a mouse metallothionein on the cell surface of Ralstonia eutropha CH34 for immobilization of heavy metals in soil. Nat. Biotechnol. 18: 661–665 Van der Lelie D, Schwitzguebel JP, Glass DJ & Baker A (2001) Assessing phytoremediation’s progress in the United States and Europe. Environ. Sci. Technol. 35(21), 447A–452A Van Hullebusch ED, Zandvoort MH & Lens PNL (2003) Metal immobilisation by biofilms : Mechanisms and analytical tools. Re/Views Environ. Sci. Bio/Technol. 2: 9–33

156 Van Hullebusch ED, Lens PNL & Tabak HH (2005) Developments in bioremediation of soils and sediments polluted with metals and radionuclides. 3. Influence of Chemical Speciation and bioavailability on contaminants immobilization/mobilization bio-processes. Re/Views in Environmental Science & Bio/Technology, 4: 185–212 Vangronsveld J (2000) In-Situ inactivation and phytoremediation of metal and metalloid contaminated soils In: Wise J, Cichon E, Inyang H & Stotmeister U (Eds.) Bioremediation of Contaminated Soils (pp 859–894). Marcel Dekker Inc., New York Van Roy S, Peys K, Dresselaers T & Diels L (1997) The use of an Alcaligenes eutrophus biofilm in a membrane bioreactor for heavy metal recovery. Res. Microbiol. 148: 526–528 Volesky B & Holan ZR (1995) Biosorption of heavy metals. Biotechnol. Prog. 11: 235–250 Wagner-Do¨bler I (2003) Pilot plant for bioremediation of mercury-containing industrial wastewater. Appl. Microbiol. Biotechnol 62: 124–133 Wang CL, Clark DS & Keasling JD (2001) Analysis of an engineered sulfate reduction pathway and cadmium precipitation on the cell surface. Biotechnol. Bioeng. 75: 285–291 Wang P, Mori T, Toda K & Ohtake H (1990) Membraneassociated chromate activity from Enterobacter cloacae. J. Bacteriol. 172: 1670–1672 Wang X & Brusseau ML (1993) Solubilization of some lowpolarity organic compounds by hydroxypropyl-Beta-cyclodextrin. Environ. Sci. Technol. 27: 2821–2825 Wang X & Brusseau ML (1995) Simultaneous complexation of organic compounds and heavy metals by a modified cyclodextrin. Environ. Sci. Technol. 29: 2632–2635 Warren LA & Haack EA (2001) Biogeochemical controls on metal behaviour in freshwater environments. Earth Sci. Rev. 54: 261–320 Warren LA, Maurice PA, Parmar N & Ferris FG (2001) Microbially mediated calcium carbonate precipitation: Implications for interpreting calcite precipitation and for solidphase capture of inorganic contaminants. Geomicrobiol. J. 18: 93–115 Watanabe ME (1997) Phytoremediation on the brink of commercialization. Environ. Sci . Technol. 31(4), 182A–186A Watson JHP & Ellwood DC (1988) Presented at the Proc. Int. Conf on Control of Environmental Problems from Metal Mines, Roros, Norway

Watson JHP & Ellwood DC (1994) Biomagnetic separation and extraction process for heavy metals from solution. Minerals Eng. 7: 1017–1028 Weidemann DP, Tanner RD, Strandberg GW & Shumate SE (1981) Modelling the rate of transfer of uranyl ions onto microbial cells. Enzyme Microbial. Technol. 3: 33–40 Whicker FW, Hinton TG, MacDonell MM, Pinder JE & Habegger LJ (2004) Avoiding destructive remediation at DOE sites. Science 303: 1615–1616 White C & Gadd GM (1998) Accumulation and effects of cadmium on sulfate reducing bacterial biofilms. Microbiology 144: 1407–1415 White C, Dennis JS & Gadd GM (2003) A mathematical process model for cadmium precipitation by sulfate-reducing bacterial biofilms. Biodegradation 14: 139–151 White C, Sharman AK & Gadd GM (1998) An integrated microbial process for the bioremediation of soil contaminated with toxic metals. Nat. Biotechnol. 16: 572–575 White C, Wilkinson SC & Gadd GM (1995) The role of microorganisms in biosorption of toxic metals and radionuclides. Int. Biodeter. Biodegrad. 35: 17–40 White C & Gadd GM (1987) Inhibition of H+ efflux and K+ uptake and induction of K+ efflux in yeast by heavy metals. Tox. Assess. 2: 437–447 White C & Gadd GM (2000) Copper accumulation by sulfatereducing bacterial biofilms. FEMS Microbiol. Lett. 183: 313–318 Wu L, Mantgem PJvan & Guo X (1996) Effects of forage plant and field legume species on soil selenium redistribution, leaching and bioextraction in soils contaminated by agricultural drain water sediment. Arch. Environ. Contam. Toxicol. 31: 329–338 Yong P & Macaskie LE (1995) Enhancement of uranium bioaccumulation by a Citrobacter sp. via enzymaticallymediated growth of polycrystalline NH4UO2PO4. J. Chem. Technol. Biotechnol. 63: 101–108 Zehnder AJB & Stumm W (1988). Geochemistry and biogeochemistry of anaerobic habitats. In: Zehnder AJB (Ed) Anaerobic Microbiol (pp 1–38). Wiley, New York Zorpette G (1996). Hanford’s nuclear wasteland. Sci. Am. May: 88–97