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Chapter 18. Geomicrobial Interactions with Chromium, Molybdenum, Vanadium, Uranium, Polonium, and Plutonium

Chapter 18. Geomicrobial Interactions with Chromium, Molybdenum, Vanadium, Uranium, Polonium, and Plutonium

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1978; Bopp et al., 1983; Cervantes, 1991; Silver and Walderhaug, 1992). In Pseudomonas fluorescens LB300, resistance to chromate (CrO42−) was found to be due to decreased chromate uptake

(Bopp, 1980; Ohtake et al., 1987). In P. ambigua, chromate resistance was attributed to formation

of a thickened cell envelope that reduced permeability of Cr(VI) and to an ability to reduce Cr(VI)

to Cr(III) (Horitsu et al., 1987).

The basis for resistance to dichromate (Cr2O72−) and chromite (Cr3+) has not been clearly established. The resistance mechanism for dichromate need not be the same as for chromate, because

Pseudomonas fluorescens LB300, which is resistant to chromate, is much more sensitive to dichromate (Bopp, 1980; Bopp et al., 1983). Some bacteria have an ability to accumulate chromium. In at

least some cases, the accumulation may be due to adsorption (Coleman and Paran, 1983; Johnson

et al., 1981; Marques et al., 1982).



Acidithiobacillus thiooxidans and A. ferrooxidans have been found to solubilize only a limited

amount of chromium contained in the mineral chromite (Cr2O3) with sulfuric acid generated by the

oxidation of sulfur (Ehrlich, 1983). Similarly, acid produced during iron oxidation by A. ferrooxidans was able to solubilize only limited amounts of chromium from chromite (Wong et al., 1982).

On the contrary, chromium can be successfully leached from some solid industrial wastes with

biologically formed sulfuric acid (Bosecker, 1986).



No observations of enzymatic oxidation of Cr(III) to Cr(VI) have been reported. However, nonenzymatic oxidation of Cr(III) to Cr(VI) may occur in soil environments, where biogenic (or abiogenic) Mn(III) or Mn(IV)- oxides may oxidize Cr(III) to Cr(VI) (Bartlett and James, 1979). These

interactions can be summarized as

Cr 3ϩ ϩ 3ΜnΟΟΗ ϩ Ηϩ → CrO24Ϫ ϩ 3Μn 2ϩ ϩ 2Η 2Ο


2Cr 3ϩ ϩ 3ΜnΟ 2 ϩ 2Η 2Ο → 2CrO24Ϫ ϩ 3Μn 2ϩ ϩ 4Ηϩ


Such oxidation can be detrimental if the Cr(VI) produced reaches a toxic level. Similar observations

were made by Chen et al. (1997) and by Kozuh et al. (2000). The latter emphasized that Cr(III) oxidation by Mn(IV) is favored by low organic matter concentration and high concentration of Mn(IV)




A number of bacterial species have been shown to reduce Cr(VI) to Cr(III) (Romanenko and

Koren’kov, 1977; Horitsu et al., 1978; Lebedeva and Lyalikova, 1979; Shimada, 1979; Kvasnikov

et al., 1985; Gvozdyak et al., 1986; Wang et al., 1989; Ishibashi et al., 1990; Shen and Wang, 1993;

Llovera et al., 1993; Lovley and Phillips, 1994; Gopalan and Veeramani, 1994; Garbisu et al., 1998;

Philip et al., 1998; Wani et al., 2007). They include Achromobacter eurydice, Aeromonas dechromatica, Agrobacterium radiobacterr strain EPS-916, Arthrobacterr spp., Bacillus subtilis, B. cereus,

B. coagulans, Desulfovibrio vulgaris (Hildenborough) ATCC 29579, Escherichia coli K-12 and

ATCC 33456, Enterobacter cloacae HO1, Flavobacterium devorans, Sarcina flava, Micrococcus

s MR-1, and Burkholderia

roseus, Pseudomonas spp., Shewanella putrefaciens (now Sh. oneidensis)

cepacia MCMB-821. It is unclear whether all these strains reduce Cr(VI) enzymatically. Sulfatereducing bacteria can reduce chromate with the H2S they produce from sulfate (Bopp, 1980), but

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Geomicrobial Interactions with Cr, Mo, V, U, Po, and Pu


D. vulgaris can do it enzymatically as well (Lovley and Phillips, 1994). Arias and Tebo (2003)

observed, however, that sulfate-reducing bacteria in general are inhibited at elevated Cr(VI) concentrations so that chromate reduction by biogenic H2S is likely to be significant only at low Cr(VI)

concentrations in the environment. T. ferrooxidans can reduce dichromate with partially reduced

sulfur species it forms during the oxidation of elemental sulfur (Sisti et al., 1996). Of those bacteria that reduce Cr(VI) enzymatically, some of the facultative strains reduce it only anaerobically,

whereas others will do it aerobically and anaerobically. Many bacterial strains reduce Cr(VI) as a

form of respiration, but at least one (P.

( ambigua G-1) reduces it as a means of detoxification (Horitsu

et al., 1987). Kwak et al. (2003) showed that the Cr(VI) reductase of P. ambigua G-1 is homologous

with the nitroreductase of strains KCTC of Vibrio harveyi and strain DH5α of E. coli. Marsh et al.

(2000) explored some of the factors that affect biological chromate reduction in microcosms of

sandy aquifer material. They found that biological reduction occurred only in light-colored sediment, abiotic reduction being observed in black, claylike sediment. The pH optimum for Cr(VI)

reduction in this sediment was 6.8. Although they detected two temperature optima, at 22 and 50°C,

the lower optimum probably represented that of the dominant bacterial group in view of the ambient in situ temperature of the sediment. The presence of oxygen in the sediment was inhibitory as

was the addition of nitrate, but not of selenate or ferrous iron. The presence of Cr(VI) prevented a

loss of sulfate or production of Fe(II). Molybdate, an inhibitor of sulfate reduction, inhibited Cr(VI)

reduction only at concentrations 40 times that of Cr(VI). Bromoethanesulfonic acid, an inhibitor

of methanogenesis, strongly inhibited Cr(VI) reduction at 20 mM concentration, but only slightly

inhibited it at concentrations between 0.2 and 2.0 mM.

Pseudomonas fluorescens LB 300, isolated from the upper Hudson River (New York State), can

reduce chromate aerobically with glucose or citrate as electron donor (Bopp and Ehrlich, 1988;

DeLeo and Ehrlich, 1994). Conditions under which aerobic reduction has been studied include batch

culturing with shaking at 200 rpm and continuous culturing with stirring and forced aeration (DeLeo

and Ehrlich, 1994). The organism converts chromate to Cr3+ in batch culture when growing in glucose–mineral salts solution (Vogel–Bonner [VB] medium) and in continuous culture (chemostat)

when growing in a citrate–yeast extract–tryptone solution buffered with phosphate (Figure 18.1).

Anaerobically, P. fluorescens strain LB300 was found to reduce chromate only when growing with

acetate as energy source (electron donor). Furthermore, although P. fluorescens LB300 will reduce

chromate aerobically at an initial concentration as high as 314 µg mL−1, anaerobically it reduces

chromate only at a concentration below 50 µg mL−1 (Bopp and Ehrlich, 1988; DeLeo and Ehrlich,

1994). Other bacteria that can reduce chromate aerobically and anaerobically include Escherichia

coli ATCC 33456 (Shen and Wang, 1993), Agrobacterium radiobacterr EPS-916 (Llovera et al.,

1993), and Burkholderia cepacia MCMB-821 (Wani et al., 2007). Reduction of chromate by E. coli

ATCC 33456 is, however, partially repressed by oxygen through uncompetitive inhibition (Shen and

Wang, 1993). Reduction of chromate by resting cells of A. radiobacterr EPS-916 proceeded initially

at similar rates aerobically and anaerobically but subsequently slowed significantly in air (Llovera

et al., 1993). P. putida PRS2000 reduces chromate aerobically more rapidly than anaerobically

(Ishibashi et al., 1990). Pseudomonas sp. strain C7 has so far been tested only aerobically (Gopalan

and Veeramani, 1994). Burholderia cepacia MCMB-821 reduces chromate equally efficiently aerobically and anaerobically with lactose as electron donor at pH 9—the organism being an alkaliphile

(Wani et al., 2007).

By contrast, Pseudomonas dechromaticans, P. chromatophila, Enterobacter cloacae OH1, and

Desulfovibrio vulgaris reduce Cr(VI) only anaerobically with organic electron donors, or H2 in

the case of D. vulgaris (Romanenko and Koren’kov, 1977; Lebedeva and Lyalikova, 1979; Komori

et al., 1989; Lovley and Phillips, 1994). Except for E. cloacae OH1, these organisms cannot use glucose as reductant. P. dechromatican and P. dechromatophila appear to be able to reduce chromate

and dichromate.

Cell extracts of Pseudomonas fluorescens LB300 reduce chromate with added glucose or NADH

(Figure 18.2). One or more plasma membrane components appear to be required (Bopp and Ehrlich,

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1988). Enterobacter cloacae HO1 also uses a membrane-bound respiratory system to reduce chromate, but it functions only under anaerobic conditions (Wang et al., 1991). By contrast, most of the

chromate-reducing activity in Escherichia coli ATCC 33456 appears to be soluble, that is, it does

not involve plasma membrane components, but is mediated by NADH (Shen and Wang, 1993). The

chromate-reducing activity of P. putida PRS2000 also does not depend on plasma membrane components. It mediates reduction via NADH or NADPH (Ishibashi et al., 1990). Desulfovibrio vulgaris

ATCC 29759 uses its cytochrome c3 as its Cr(VI) reductase coupled to hydrogenase when using H2

as reductant (Lovley and Phillips, 1994). Ackerley et al. (2004) detected soluble Cr(VI)-reducing

activity in aerobically growing strains of P. putida and E. coli associated with soluble bacterial flavoproteins. The flavoprotein from P. putida, labeled ChrR, and that from E. coli, labeled YieF, were

dimers. They tested the activity of the two reductases with NADH as electron donor. Unlike the

YieF dimer, the ChrR dimer generated a flavin semiquinone transiently and reactive oxygen species




A382 nm


















FIGURE 18.1 Chromate reduction by resting and growing cells of chromate-resistant Pseudomonas fluorescens LB300. (A) Resting cells grown with and without chromate. (△) Chromate-grown cells in the absence of

electron donor (results were the same for cells grown without chromate and assayed in the absence of electron

donor; no chromate reduction was observed; (●) chromate-grown cells with 0.5% (w/v) glucose; (▲) cells

grown without chromate and assayed with 0.5% (w/v) glucose. Chromate was not reduced by spent medium

from either chromate-grown cells or cells grown without chromate or by assay buffer containing either 0.25 or

0.5% glucose. Chromate concentration was measured as absorbance at 328 nm after cell removal. (B) Growing

cells in VB broth at an initial K 2CrO4 concentration of 40 µg mL−1; growth of the culture was measured

photometrically as turbidity at 600 nm; chromate concentration was measured as absorbance at 382 nm after

first removing cells from replicate samples by centrifugation followed by filtration. ((A) and (B): From Bopp

LH, Ehrlich HL, Arch. Microbiol., 150, 426–431, 1988. With permission of Springer Science and Business

media.) (C) Chromate reduction by cells growing in citrate-chromate medium in a chemostat at a dilution rate

of 1.17 mL h−1; (△) chromate concentration in uninoculated reactor; (○) chromate concentration in inoculated

reactor; (□) cell concentration in inoculated reactor. ((C): From DeLeo PC, Ehrlich HL, Appl. Microbiol.

Biotechnol., 40, 756–759, 1994. With permission of Springer Science and Business media.)

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Hexavalent chromium (mg


Cell concentration (mg L−1)






Time (h)



Absorbance (600 nm)

Absorbance (382 nm)

Geomicrobial Interactions with Cr, Mo, V, U, Po, and Pu


diminished over time, which suggested that Cr(V) may be an intermediate in Cr(VI) reduction by

this reductase. Because the YieF dimer did not generate flavin semiquinone and used only 25% of

the electrons from NADH for reduction of reactive oxygen species, Ackerley et al. (2004) suggest

that in reducing Cr(VI), this reductase may transfer three electrons to Cr(VI) and one to reactive

oxygen species for every two NADH oxidized.

In anaerobically grown Shewanella putrefaciens (now Sh. oneidensis)

s MR-1, chromate reductase

activity is associated with the cytoplasmic membrane (Myers et al., 2000). Although this organism

is facultative, it reduces chromate only anaerobically. Both formate and NADH but not l-lactate or

NADPH can serve as electron donors to the Cr(VI) reductase system, which includes a multicomponent electron transport system. Some of the activity is inducible in cells grown anaerobically with

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A382 nm














FIGURE 18.2 Chromate reduction by cell extract from Pseudomonas fluorescens LB300. GL

L−, without

added glucose; GL

L+, with glucose from time 0 and KCN and NaN3 at time indicated. Chromate concentration

was measured as absorbance at 382 nm after cell removal. (From Bopp LH, Ehrlich HL, Arch. Microbiol.,

150, 426–431, 1988. With permission of Springer Science and Business Media.)

lactate as electron donor and nitrate or fumarate as electron acceptor, but this activity is inhibited by

nitrite (Viamajala et al., 2002). Although Cr(III) is the usual product of respiratory Cr(VI) reduction by Sh. oneidensis and other unrelated Cr(VI) reducers, one case in which Cr(II) accompanied

Cr(III) in Cr(VI) respiration by Sh. oneidensis MR-1 has been reported (Daulton et al., 2007). The

Cr(II) in this instance was concentrated near the cytoplasmic membrane.

18.1.6 IN SITU


In situ rates of microbial Cr(VI)-reducing activity are generally not readily available in the literature, although such measurements would be useful because a number of different bacteria possess

the ability to reduce Cr(VI). On the contrary, Wang and Shen (1997) examined rate parameters for

a number of pure cultures under laboratory conditions. For example, they reported that the halfmaximal Cr(VI) reducing velocity constant Ks, in mg Cr(VI) L−1, is 5.43 for Bacillus subtilis (aerobic),

19.2 for Desulfovibrio vulgaris ATCC 29579 (anaerobic), 8.64 for Escherichia coli ATCC 33456

(anaerobic), 641.9 for Pseudomonas ambigua G-1 (aerobic), and 5.55 for P. fluorescens LB300 (aerobic), based on Monod kinetics. Natural levels of Cr(VI) in most environments can be expected to

be low. However, anthropogenic pollution can cause very significant elevation in environmental

chromium concentrations.

Lebedeva and Lyalikova (1979) isolated a strain of Pseudomonas chromatophila from the effluent of a chromite mine in Yugoslavia that contained the mineral crocoite (PbCrO4) in its oxidation

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Geomicrobial Interactions with Cr, Mo, V, U, Po, and Pu


zone. At this site, the organism clearly found a plentiful source of Cr(VI). The isolate was shown

to use a range of carbon and energy sources anaerobically for chromate reduction. These included

ribose, fructose, benzoate, lactate, acetate, succinate, butyrate, glycerol, and ethylene glycol but not

glucose or hydrogen. Crocoite reduction with lactate as electron donor was demonstrated anaerobically in the laboratory.



The first practical application of microbial Cr(VI) reduction as a bioremediation process was explored

by Russian investigators. They presented evidence that indicated that bacterial chromate reduction

can be harnessed in wastewater and sewage treatment to remove chromate (Romanenko et al., 1976;

Pleshakov et al., 1981; Serpokrylov et al., 1985; Simonova et al. 1985). The process also has potential

application in treatment of tannery and, especially, electroplating wastes and in situ bioremediation.

In the case of tannery and electroplating waste treatment, prior dilution of the waste may be necessary to bring the Cr(VI) concentration into a range tolerated by the Cr(VI)-reducing bacteria.

Extensive research in the use of Enterobacter cloacae HO1 in bioremediation of Cr(VI)containing wastewaters was performed in Japan (Ohtake et al., 1990; Komori et al., 1990a,b;

Yamamoto et al., 1993; Fuji et al., 1994).




Molybdenum is an element of the second transition series in the periodic table. In mineral form,

it occurs extensively as molybdenite (MoS2 ). The minerals wulfenite (PbMoO4 ) and powellite

(CaMoO4 ) are often associated with the oxidation zone of molybdenite deposits (Holliday, 1965).

Molybdite (MoO3) is another molybdenum-containing mineral that may be encountered in nature.

The abundance of Mo has been reported to be 2–4 g t−1 in basaltic rock, 2.3 g t−1 in granitic rock,

and 0.001–0.005 g t−1 in ocean waters (Enzmann, 1972).

The oxidation states in which molybdenum can exist include 0, +2, +3, +4, +5, and +6. Of

these, the +4 and +6 states are the most common, but the +5 state is of biological significance.

Molybdenum oxyanions of the +6 oxidation state tend to polymerize the complexity of the polymers depending on the pH of the solution (Latimer and Hildebrand, 1942).

Molybdenum is a biologically important trace element. A number of enzymes feature it in their

structure, for example, nitrogenase, nitrate reductase (Brock and Madigan, 1991), sulfite reductase,

and arsenite oxidase (Anderson et al., 1992). Molybdate is an effective inhibitor of bacterial sulfate

reduction (Oremland and Capone, 1988).



Molybdenite (MoS2) is aerobically oxidizable as an energy source by Acidianus brierleyi (Brierley

and Murr, 1973) (see also Chapter 20) with the formation of molybdate and sulfate. Acidithiobacillus

ferrooxidans can also oxidize molybdenite, but it is poisoned by the resulting molybdate (Tuovinen

et al., 1971) unless the molybdate is rendered insoluble, for instance by reaction with Fe3+. Sugio

et al. (1992) reported that Thiobacillus (now Acidithiobacillus)

s ferrooxidans AP-19-3 contains an

enzyme that oxidizes molybdenum blue (Mo5+) to molybdate (Mo6+). They purified the molybdenum oxidase and found it to be an enzyme complex that included cytochrome oxidase as an important component. The function of molybdenum oxidase in the organism remains unclear in view of

its sensitivity to molybdate.

Molybdate was first shown to be reduced by Sulfolobus sp. by Brierley and Brierley (1982).

In a more detailed study, molybdate was shown to be reduced microbially to molybdenum blue

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(containing Mo5+) by Acidithiobacillus ferrooxidans using sulfur as electron donor (Sugio et al.,

1988). The enzyme that reduced the Mo6+ was identified as sulfur:ferric ion oxidoreductase.

Molybdate has also been shown to be reduced anaerobically to molybdenum blue by Enterobacter

cloacae strain 48 using glucose as electron donor (Ghani et al., 1993). The reduction appears to be

mediated via NAD and b-type cytochrome. Other bacteria reported to be able to reduce Mo(VI)

include Pseudomonas guillermondii, Micrococcus sp., and Desulfovibrio desulfuricans (reviewed

by Lloyd, 2003).



Vanadium belongs to the first transition series of elements in the periodic table. In mineral form, it

often occurs in complex forms such as patronite (a complex sulfide), roscoelite (a vanadium mica),

vanadinite (a lead vanadate), and carnotite (a hydrous potassium uranium vanadate) (DeHuff, 1965).

Its average abundance (milligrams per kilogram) in granites is 72, in basalts 270, and in soil 90. Its

average concentration (nanograms per cubic meter) in freshwater is 0.0005 and in seawater 0.0025

(Bowen, 1979).

Vanadium occurs in the oxidation states of 0, +2, +3, +4, and +5. Pentavalent vanadium in

solution occurs as the oxyanion VO3− (vanadate) and is colorless. Tetravalent vanadium in solution

occurs as VO2+ and is deep blue. Trivalent vanadium (V3+) forms a green solution, and divalent

vanadium (V2+) a violet solution (Dickerson et al., 1979).

As a trace element in prokaryotes, vanadium has been found to occur in place of molybdenum

in certain nitrogenases (Brock and Madigan, 1991) (see also Chapter 13). It also occurs in oxygencarrying blood pigment of ascidian worms.



Five different bacteria have been reported to be able to reduce vanadate. The first three are Veillonella



s) lactilyticus, Desulfovibrio desulfuricans, and Clostridium pasteurianum, which

were shown by Woolfolk and Whiteley (1962) to be able to reduce vanadate to vanadyl with hydrogen using a hydrogenase,

VOϪ3 ϩ Η 2 → VO(OH) ϩ OHϪ


The fourth and fifth organisms are new isolates assigned to the genus Pseudomonas (Yurkova

and Lyalikova, 1990; Lyalikova and Yurkova, 1992). One of these, isolated from a waste stream

from a ferrovanadium factory, was named P. vanadiumreductans, and the other, isolated from

seawater in Kraternaya Bay, Kuril Islands, P. issachenkovii. Both are gram-negative, motile,

non-spore-forming rods that can grow as facultative chemolithotrophs and facultative anaerobes.

Anaerobically, chemolithotrophic growth was observed with H2 and CO as alternative energy

sources, CO2 as carbon source, and vanadate as terminal electron acceptor. However, the organisms can also grow organotrophically under anaerobic conditions with glucose, maltose, ribose,

galactose, lactose, arabinose, lactate, proline, histidine, threonine, and serine as carbon and energy

sources. P. issachenkovii can also use asparagine as carbon and energy source. Vanadate reduction by the organism involved transformation of pentavalent vanadium to tetravalent and trivalent

vanadium. The tetravalent oxidation state was identified in the medium by development of a blue

color and the trivalent state by formation of a black precipitate and by its reaction with tairon

reagent. An equation describing the overall reduction of vanadium by lactate in these experiments

was presented by the authors as

2NaVO3 + NaC3H5O3 → V2O3 + NaC2H3O2 + NaHCO3 + NaOH

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Geomicrobial Interactions with Cr, Mo, V, U, Po, and Pu


It accounts for the alkaline pH developed by the medium during growth that started at pH 7.2.

Antipov et al. (2000) found that molybdenum- and molybdenum cofactor-free nitrate reductases of

P. issachenkovii appear to mediate the vanadate reduction. Homogeneous membrane-bound nitrate

reductase from the organism reduced vanadate with NADH as electron donor. In a medium containing both nitrate and vanadate, the organism reduced nitrate before vanadate. A vanadium mineral

similar to sherwoodite was detected in cultures reducing vanadium, suggesting that these bacteria

may play a role in epigenetic vanadium mineral formation (Lyalikova and Yurkova, 1992).

More recently, Shewanella oneidensis MR-1 was shown to be able to use vanadate as sole electron acceptor anaerobically, that is, it was able to respire anaerobically on lactate, pyruvate, formate,

fumarate, Fe(III), or citrate using vanadate as terminal electron acceptor (Carpentier et al., 2003,

2005). It reduces vanadate [VV] to vanadyl [VIV] (Carpentier et al., 2003). The reduction was inhibited by 2-heptyl-4-hyroxyquinoline N

N-oxide and antimycin A, and proton translocation associated

with vanadate reduction was abolished by the protonophores dinitrophenol and carbonyl cyanide

m-chlorophenuylhydrazone (Carpentier et al., 2005).

Geobacter metallireducens has also been shown to be able to grow anaerobically with acetate as

electron donor and vanadate as terminal electron acceptor (Ortiz-Bernad et al., 2004a). It reduced

VV to VIV, which subsequently precipitated. Microprobe analysis of the precipitate suggested that it

could have been vanadyl phosphate. In a bioremediation study of groundwater contaminated with

uranium(VI) and vanadium(V), acetate stimulation of Geobacteraceae caused precipitation of the

vanadium as a result of its reduction (Ortiz-Bernad et al., 2004a).

Vanadium(V) can also be reduced nonenzymatically by bacteria. An example is the reduction

of vanadium(V), at concentrations up to 5 mM, to vanadium(IV) by Acidithiobacillus thiooxidans,

using elemental sulfur as its energy source. The vanadium reduction in this instance is brought

about by partially reduced sulfur intermediates produced by the organism during oxidation of the

elemental sulfur (Briand et al., 1996).




Uranium is one of the naturally occurring radioactive elements. Its abundance in the Earth’s crust

is only 0.0002%. It is found in more than 150 minerals, but the most important are the igneous

minerals pitchblende and coffinite and the secondary mineral carnotite (Baroch, 1965). It is found

in small amounts in granitic rocks (4.4 mg kg−1) and in even smaller amounts in basalt (0.43 mg kg−1).

In freshwater it has been reported in concentrations of 0.0004 mg kg−1 and in seawater, 0.0032 mg kg−1

(Bowen, 1979).

Uranium can exist in the oxidation states 0, +3, +4, +5, and +6 (Weast and Astle, 1982). The

+4 and +6 oxidation states are of greatest significance microbiologically. In nature, the +4 oxidation state usually manifests itself in insoluble forms of uranium, for example, UO2. The +6 oxidation state predominates in nature in soluble, and hence mobile, form, for example, UO22+ (Haglund,

1972). In radioactive decay of an isotopic uranium mixture, alpha, beta, and gamma radiation are

emitted, but the overall rate of decay is very slow because the dominant isotopes have very long

half-lives (Stecher, 1960). This slow rate of decay probably accounts for the ability of bacteria to

interact with uranium species without experiencing lethal radiation damage.



Acidithiobacillus ferrooxidans has been shown to oxidize tetravalent U4+ to hexavalent UO22+ in a

reaction that yields enough energy to enable the organism to fix CO2. Nevertheless, experimental

demonstration of growth of A. ferrooxidans with U4+ as sole energy source has not succeeded to

date. Thiobacillus acidophilus (now renamed Acidiphilium acidophilum)

m was also found to oxidize

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U4+ but without energy conservation (DiSpirito and Tuovinen, 1981, 1982a,b) (see also Chapter 20).

Recently, the autotroph, Thiobacillus denitrificans, was shown to promote oxidation of U(IV) oxide,

uraninite, anaerobically in the presence of nitrate in an in-laboratory glove-box experiment (Beller,

2005). Although U(IV) oxidation was accompanied by significant nitrate reduction, the amount of

nitrate consumed significantly exceeded the amount of U(IV) oxidized. The excess nitrate reduction

was attributable to biooxidation of H2 in the atmosphere in the glove box (Beller, 2005).



A number of organisms have been shown to be able to reduce hexavalent uranium (UO22+) to tetravalent uranium (UO2). The first demonstration was with Veilonella (Micrococcus


s) lactilyticus

using H2 as electron donor under anaerobic conditions (Woolfolk and Whiteley, 1962). Much more

recently, some other bacteria were shown to be able to reduce U(VI) to U(IV) anaerobically. They

are the facultative organism Shewanella putrefaciens (Lovley et al., 1991) and Sh. alga strain BrY

(Caccavo et al., 1992) and the strict anaerobe Geobacter metallireducens strain GS-15 (Lovley

et al., 1991, 1993a), Desulfovibrio desulfuricans (Lovley and Phillips, 1992), D. vulgaris (Lovley et al.,

1993b), and Desulfovibrio sp. (Pietzsch et al., 1999). In Desulfovibrio, cytochrome c3 appears to be

the U(VI) reductase (Figure 18.3) (Lovley et al., 1993b). The electron donors used by these organisms may be organic or, in some instances, H2. Although G. metallireducens and Sh. putrefaciens

can gain energy from the process of U(VI) reduction (Lovley et al., 1991), D. desulfuricans and

D. vulgaris are unable to do so (Lovley and Phillips, 1992; Lovley et al., 1993c). Desulfovibrio sp.,

on the contrary, is able to gain energy from the process (Pietzsch et al., 1999).



U(VI) (mM)


Cytochrome, hydrogenases, hydrogen

Cytochrome, hydrogenases, no hydrogen

Cytochrome, hydrogen

Hydrogenases, hydrogen

Cytochrome, periplasmic hydrogenase, hydrogen











FIGURE 18.3 Reduction of U(VI) by electron transfer from H2 to U(VI) via hydrogenase and cytochrome c3.

As noted, pure periplasmic hydrogenase or a protein fraction containing two hydrogenases was used. (From

Lovley DR, Widman PK, Woodward JC, Phillips JEP, Appl. Environ. Microbiol., 59, 3572–3576, 1993b. With


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Geomicrobial Interactions with Cr, Mo, V, U, Po, and Pu


Insofar as the study of electron transport pathways in U(VI) and Cr(VI) reduction under anaerobic

conditions by Shewanella oneidensis MR-1 is concerned, Bencheikh-Latmani et al. (2005), using

whole-genome DNA microarrays, showed that U(VI)-reducing conditions caused upregulation of

121 genes and Cr(VI)-reducing conditions 83 genes. Genes whose products are known to enable the

use of the alternative electron acceptors fumarate, dimethyl sulfoxide, Mn(IV), or soluble Fe(III)

were upregulated under both U(VI)- and Cr(VI)-reducing conditions. Mutant studies confirmed

that several genes, whose products are known to be involved in ferric citrate reduction, are also

involved in the reduction of both U(VI) and Cr(VI) by Sh. oneidensis MR-1. The genes are mtrA,

mtrB, mtrC, and menC. However, genes coding for efflux pumps were upregulated under Cr(VI)reducing conditions but not U(VI)-reducing conditions. These findings by Bencheikh-Latmani

et al. (2005) show that the same electron transport pathway or parts thereof may be involved in the

reduction of U(VI) and Cr(VI) and several other terminal electron acceptors.

The isolation of some of the organisms from freshwater and marine sediments suggests that this

microbial activity may play or may have played a significant role in the immobilization of uranium

and its accumulation in sedimentary rock. The first evidence in support of such immobilization

was obtained by Gorby and Lovley (1992) in experiments with groundwater amended with 0.4 or

1.0 mM uranyl acetate and 30 mM NaHCO3 and inoculated with Geobacter metallireducens GS-15,

which produced a black precipitate from the dissolved UO22+. The black precipitate was identified

as uraninite (UO2). Frederickson et al. (2000) studied U(VI) reduction by Shewanella putrefaciens

CN32 in the presence of goethite (Fe2O3 · H2O). They found that besides enzymatic reduction of the

U(VI) by Sh. putrefaciens CN32, Fe(II) that had been sorbed to goethite was able to reduce U(VI)

abiotically, as was the humic analog anthraquinone-2,6-disulfonate (AQDS).


Uranium(VI)-contaminated aquifer groundwater can be bioremediated by reduction of the U(VI)

to insoluble U(IV) under various conditions. Thus in a field experiment, the uranium level in a

contaminated aquifer in Rifle, Colorado, United States, was lowered to <0.18 µM from a range of

0.4 to 1.4 µM within 50 days by injection of 1–3 mM acetate per day (Anderson et al., 2003). Initial

loss of uranium from the groundwater was attributed to the activity of Geobacterr spp., which can

reduce U(VI) enzymatically as well as by producing Fe(II) from reduction of Fe(III). But as acetate

injection continued, Geobacterr spp. was gradually replaced by sulfate reducers, stimulated by the

injected acetate, which they used as carbon and energy source, and by using sulfate in the groundwater as terminal electron acceptor, and causing the uranium concentration in the groundwater to

increase because the particular sulfate reducers that became selectively enriched were incapable of

reducing U(VI) or promoting its reduction. It appears that sustained bioremediation of a uraniumcontaminated aquifer by Geobacterr spp. would benefit from maintenance of conditions that favor

the activity of Geobacterr spp. over sulfate-reducers. It also appears that only U(VI) dissolved in the

groundwater at the Rifle, Colorado, aquifer site is amenable to bioremediation. U(VI) bound to the

sediment in the aquifer was not microbially reducible (Ortiz-Bernad et al., 2004b).

Suzuki et al. (2005) have presented evidence from shallow water sediment from an open pit of a

uranium mine in Washington State, United States, that showed U(VI) immobilization by bioreduction to U(IV) oxide, in which both Fe(III)- and sulfate-reducers, members of the Geobacteraceae

and Desulfovibrionacea, respectively, were implicated. Both families of organisms were previously

known to contain members capable of U(VI) reduction.

Nevin et al. (2003) demonstrated uranium bioremediation in a high-salinity subsurface sediment

from an aquifer associated with uranium mine tailings at Shiprock, New Mexico, United States, by

stimulation with acetate, which enriched populations of microorganisms related to Pseudomonas

and Desulfosporosinus species.

Shelobolina et al. (2004) obtained evidence for potential U(VI) anaerobic bioreduction at low

pH in nitrate- and U(VI)-contaminated subsurface sediment from the Natural and Accelerated

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Bioremediation Research Field Center in Oak Ridge, Tennessee, United States. Salmonella subterranean sp. nov., a gram-negative, motile rod capable of using O2, NO3−, S2O32−, fumarate, and

malate as terminal electron acceptors and of reducing U(VI) in cell suspension was isolated from

the sediment.



Polonium is a radioactive element that occurs naturally in association with uranium and thorium

minerals. Different isotopes of polonium are produced in the decay of 238U, 235U, and 232Th. Of

these isotopes, 210Po, which originates from the decay of 238U, has the longest half-life (138.4 days)

(Lietzke, 1972). Its known oxidation states are +2 and +4 (see LaRock et al., 1996).

Polonium-210 can be an environmental pollutant arising by release from uranium-containing

phosphorite, which is commercially exploited for its phosphate and phosphogypsum, which is a byproduct in the manufacture of phosphoric acid from phosphorite (see LaRock et al., 1996).

Sulfate-reducing bacteria that were found by LaRock et al. (1996) to be associated with phosphogypsum are able to mobilize Po contained in the phosphogypsum. In laboratory experiments, this

mobilization required that the dissolved sulfate concentration in the bulk phase was below 10 µM.

Above this sulfate concentration, enough H2S was produced to coprecipitate the mobilized Po as a

metal sulfide (LaRock et al., 1996). The release of polonium apparently depended on the reduction

of the sulfate in the gypsum. Aerobic bacteria were also able to mobilize Po in the phosphogypsum.

The mechanism in this instance may involve Po complexation by ligands produced by the active

organisms (LaRock et al., 1996).

Immobilization of Po by at least one aerobic bacterial isolate has also been reported (Cherrier

et al., 1995). The Po was taken into the cell by a mechanism that appeared to differ from that of

sulfate uptake. However, its partitioning after uptake paralleled that of sulfur among cell components that included cell envelope, cytoplasm, and cytoplasmic protein. Most polonium and sulfur

were detected in the cytoplasmic fraction. In nature, such immobilization of Po must be considered

transient if upon death of these cells the Po becomes redissolved in the bulk phase.



Recently enzymatic reduction of plutonium(IV) to the more soluble plutonium(III) under anaerobic

conditions was definitively demonstrated with Geobacter metallireducens GS-15 and Shewanella

oneidensis MR-1 using freshly precipitated, amorphous Pu(IV)(OH)4 and soluble Pu(IV)(EDTA)

(Boukhalfa et al., 2007). Both organisms reduced the Pu(IV)(EDTA) and amorphous Pu(IV)(OH)4

in the presence of EDTA, forming Pu(III)(EDTA). In the absence of a complexing ligand such as

EDTA, Sh. oneidensis MR-1 produced only minor amounts of Pu(III) and G. metallireducens produced little or no Pu(III). The Pu(IV) reduction did not support growth of either organism.

The only previous observation suggesting bacterial interaction with plutonium involved a study

with Bacillus polymyxa and B. circulans (Rusin et al., 1994). However, reduction of Pu(IV) to

Pu(III) by the bacteria was only inferred but not directly demonstrated.



Enzymatic oxidation of Cr(III) by bacteria has not been demonstrated. Nonenzymatic oxidation of

Cr(III), which is dependent on biogenic (bacterial, fungal) formation of Mn(IV), which then oxidizes Cr(III) to Cr(VI) chemically, may occur in soil.

Aerobic and anaerobic reduction of Cr(VI) by bacteria has been demonstrated. The process is

in many instances a form of respiration. Various organic electron donors may serve, but not all act

equally well aerobically and anaerobically. Chromate and dichromate are not necessarily reduced

equally well. The ability to reduce chromate does not always correlate with chromate tolerance.

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Chapter 18. Geomicrobial Interactions with Chromium, Molybdenum, Vanadium, Uranium, Polonium, and Plutonium

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