Bacterial Pu(V) reduction in the absence and presence of Fe(III)–NTA: modeling and experimental approach
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- Deo, R.P., Rittmann, B.E. & Reed, D.T. Biodegradation (2011) 22: 921. doi:10.1007/s10532-010-9451-z
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Plutonium (Pu), a key contaminant at sites associated with the manufacture of nuclear weapons and with nuclear-energy wastes, can be precipitated to “immobilized” plutonium phases in systems that promote bioreduction. Ferric iron (Fe3+) is often present in contaminated sites, and its bioreduction to ferrous iron (Fe2+) may be involved in the reduction of Pu to forms that precipitate. Alternately, Pu can be reduced directly by the bacteria. Besides Fe, contaminated sites often contain strong complexing ligands, such as nitrilotriacetic acid (NTA). We used biogeochemical modeling to interpret the experimental fate of Pu in the absence and presence of ferric iron (Fe3+) and NTA under anaerobic conditions. In all cases, Shewanella alga BrY (S. alga) reduced Pu(V)(PuO2+) to Pu(III), and experimental evidence indicates that Pu(III) precipitated as PuPO4(am). In the absence of Fe3+ and NTA, reduction of PuO2+ was directly biotic, but modeling simulations support that PuO2+ reduction in the presence of Fe3+ and NTA was due to an abiotic stepwise reduction of PuO2+ to Pu4+, followed by reduction of Pu4+ to Pu3+, both through biogenically produced Fe2+. This means that PuO2+ reduction was slowed by first having Fe3+ reduced to Fe2+. Modeling results also show that the degree of PuPO4(am) precipitation depends on the NTA concentration. While precipitation out-competes complexation when NTA is present at the same or lower concentration than Pu, excess NTA can prevent precipitation of PuPO4(am).
KeywordsPlutoniumShewanella algaBacterial reductionNTAIronBioreductionModeling
The application of biotechnology for remediation of metals and radionuclides has been well documented (Banaszak et al. 1998, 1999a, b; Caccavo et al. 1992; Farrell et al. 1999; Gorby and Lovley 1992; Haas and Dichristina 2002; Liu et al. 2002; Lovley et al. 1993; Reed et al. 2010; Rittmann et al. 2002a; Songkasiri et al. 2002; Truex et al. 1997). Unlike for organic contaminants, a bioremediation strategy for radionuclides and metals transforms them into phases that should render them immobile and recalcitrant (NRC 2000). This often proceeds via reduction of a radionuclide to an oxidation state that is less soluble.
The occurrence of plutonium (Pu) in the subsurface environment is a concern due to its long half-life (2.4 × 104 years) and toxicity (Neu et al. 2005). Under oxidizing conditions, the predominant forms of plutonium are Pu(V)O2+ and Pu(VI)O22+, which form distinct inorganic/organic complexes when present as a contaminant in groundwater (Choppin 2003; Cleveland and Rees 1981). While PuO2+ (i.e., Pu(V)) does not form hydroxyl complexes until pH > 8 and should be very mobile under most subsurface conditions, PuO22+ (i.e., Pu(VI)) forms strong hydroxyl complexes and sorbs strongly to aquifer surfaces. However, Pu(VI) is easily reduced and not likely to persist in biologically active systems unless it undergoes irreversible aggregation and polymerization, which may stabilize it against reduction, leading to enhanced subsurface mobility and persistence in the oxidized form (Francis 2007; Reed et al. 2010).
The key to plutonium immobilization in the subsurface is its reduction to Pu(IV) and Pu(III) species. Of these two reduced oxidation states, Pu(IV) exhibits a lower solubility and a much higher tendency towards aggregation, polymer formation, and sorption. From this perspective, Pu(IV) is by far the preferred target oxidation state (Francis 2007; Reed et al. 2010).
Only a few papers have been published on bioreduction of Pu under anaerobic conditions: by Bacillus polymyxa and circulans (Rusin et al. 1994), Geobacter metallireducens GS15 and Shewanella oneidensis MR1 (Boukhalfa et al. 2007; Icopini et al. 2009), and Clostridium sp. (Francis et al. 2008). Furthermore, evidence of Pu reduction by abiotic mechanisms (Rai et al. 2002; Reed et al. 2006; von Gunten and Benes 1995) suggests that bioremediation strategies in subsurface environment require an understanding of coupled abiotic and biotic processes for accurate prediction of plutonium fate in complex subsurface environment (Banaszak et al. 1999a; Lovley 1993; Neu et al. 2002; Reed et al. 2010).
Our group published work on bioreduction of higher-valent uranium and plutonium (Reed et al. 2007), where part of that work investigated Pu(V) reduction in the absence and presence of Fe3+ and NTA. Specifically, we investigated the bioreduction of Pu(V) under anaerobic conditions with Shewanella alga BrY (a facultative metal-reducing bacterium (Caccavo et al. 1992, 1996) in the presence and absence of Fe3+–NTA, an aqueous form of ferric iron. Although Pu(V) reduction occurred in both cases, reduction was significantly slower in the presence of iron. The suggested explanation was that Fe3+ was the preferred electron acceptor (over Pu(V)), but Fe2+, once generated by bioreduction, caused concurrent abiotic reduction of Pu(V).
In this study, we use modeling analyses to understand the mechanisms governing the previously observed slow reduction of Pu(V) in the presence of iron (Reed et al. 2007). We also discuss new spectroscopic results that expand our ability to interpret the experimental findings.
Materials and methods
All methods to synthesize PuO2+; assay for Pu, Fe, lactate, and other organics; determine oxidation states of PuO2+ and Fe (Fe3+ and Fe2+); grow S. alga; and carry out anaerobic experiments are the same as reported in Reed et al. (2007). For this work, we added spectroscopic analyses of the reduced Pu solid formed. The spectroscopic analysis was done using X-ray absorption near edge spectroscopy (XANES) with MR-CAT beamline at the Advanced Photon Source (APS) following the methods of Songkasiri et al. (2002).
Background information on the biogeochemical model CCBATCH
Modeling was performed using the biogeochemical model CCBATCH (Rittmann et al. 2002b; VanBriesen and Rittmann 2000a, b), which was developed by our team to quantitatively link all the different types of reactions that control the fate of radionuclides and a range of other metals and organic co-contaminants. The basic structure of CCBATCH is described here, and then special features needed to use the model to represent the experiments reported here are described. Certain details, such as parameter values, are reported as needed in “Results and discussion” section.
CCBATCH explicitly couples biological electron-donor and -acceptor consumptions to simultaneous chemical reactions in order to determine the effect of biological reactions on the fate of various components in the system. The original CCBATCH model (VanBriesen and Rittmann 1999, 2000a) couples microbially catalyzed reactions, which are kinetically controlled, with aqueous phase acid/base and complexation reactions, which are at thermodynamic equilibrium. Rittmann et al. (2002b) added a sub-model that links kinetically or equilibrium-controlled precipitation/dissolution to the microbial and aqueous-phase reactions. We used the equilibrium sub-model. CCBATCH was designed to describe batch reactions, such as those performed in this study. The microbial sub-model includes oxidation of an electron-donor substrate (e.g., lactate in our studies), synthesis and endogenous decay of biomass, stoichiometric utilization of an electron acceptor, and stoichiometric consumption or generation of inorganic carbon, ammonium–nitrogen, and acidic hydrogen. The equilibrium feature of the sub-model precipitates or dissolves just the amount of solid phase so that the aqueous phase speciation of the cation and anion match the solubility product (Ksp) for every time step. In general, precipitation consumes basic species, and the sub-model represents this through stoichiometric production of acidic hydrogen. Finally, CCBATCH computes pH based on changes in acidic hydrogen and solving a proton condition. To solve for equilibrium speciation, CCBATCH uses a Newton–Raphson technique that combines the aqueous-phase mass balances with mass action equilibrium expressions for all relevant acid/base and complexation reactions, and it can compute the equilibrium pH from the proton condition when the pH is not fixed.
Upgrading CCBATCH for anaerobic growth
(a) Complexes and formation constants for Fe3+ complexes; (b) Kinetic parameters for Fe3+ and PuO2+ reduction
(a) Complexes and formation constants for Fe3+ complexes
Smith et al. (2004)
(b) Kinetic parameters for Fe3+ and PuO2+ reduction
77.9 mole Lac. molcell−1 h−1
Best fit to data
Best fit to data
Hacherl and Kosson (2003)
0.892 M h−1
Best fit to data
26.25 M2 h−1
Best fit to data
26.25 M2 h−1
A second challenge is that the bioavailabilities of the various Fe3+ species are not known (Haas and Dichristina 2002). Since the pH of our experiments was buffered at neutral, pH has minimal effect on the relative concentrations of all Fe3+ species, and we could use total Fe3+ as the bioavailable form of Fe3+. Table 1b summarizes the parameters used to describe Fe3+ bioreduction and biomass growth from it: qlactate = maximum specific rate of lactate utilization, Ka,Fe(III) = electron acceptor concentration that gives half of the maximum growth rate, and b = first-order endogenous-decay rate. Should only one (or more than one) Fe3+ species be bioavailable, its concentration remains in a constant ratio with total Fe3+; thus, the model would represent the experimental results equally well, but with the value of qlactate increased in proportion to the concentration ratio of total Fe3+ to the bioavailable species. The bioreduction of Fe3+ was not limited by lactate, since it was in excess.
Biotic PuO2+ reduction
We added plutonium reactions into the model to represent reduction of PuO2+ in the presence and absence of Fe3+–NTA for the experiments in Reed et al. (2007). In both the cases, we chose Pu3+ as the ultimate reduced form of plutonium based on X-ray absorption near edge spectroscopy (XANES) analysis of the solids formed at the end of the experiments. The potential for Pu(V)O2+ reduction to Pu3+ has been reported before under anaerobic conditions by metal-reducing bacteria: Geobacter metallireducens GS15 and Shewanella oneidensis MR1 (Boukhalfa et al. 2007), Geobacter sulfurreducens and Shewanella oneidensis (Renshaw et al. 2009), and Clostridium sp. (Francis et al. 2008).
Formation constants for major Pu(V, IV, III) aqueous complexes at ionic strength = 0.1
Smith et al. (2004)
Banaszak et al. (1999a)
Banaszak et al. (1999a)
Log Ksp = −24.6
Adding abiotic PuO2+ reduction
Results and discussion
Figure 2B shows the model-simulated speciation of Pu3+ in the absence (solid lines) or presence (dashed lines) of Pu3+ precipitate formation, i.e., PuPO4(am). NTA is not present in the medium. When the precipitation subroutine was “turned off” (solid lines for Pu species), the major aqueous Pu3+ species formed and their respective percentages are 26.4% for Pu(OH)2+(aq) and 73.6% for Pu3+(aq). The high percentage of free Pu3+(aq) illustrates the lack of concentration and/or strength of anions in the medium to complex with all the Pu3+. “Turning on” the precipitation subroutine (dashed line) precipitated all the Pu3+ as PuPO4(am). This means that Pu3+ complexation with anions present in the medium was not strong enough to keep Pu3+ in a soluble form when NTA was not present.
Figure 4B shows the speciation of Pu3+ in the absence (i) or presence (ii) of PuPO4(am) precipitation. In both cases, almost all (99.8%) of Pu3+ was Pu3+–NTA(aq), which can be attributed to the strength of Pu3+–NTA complex (e.g., Fig. 3). The remaining ~0.2% of Pu3+ was speciated as free Pu3+(aq) and Pu(OH)2+(aq), with an additional ~0.07% of PuPO4(am) formed only when the precipitation subroutine was turned on (ii). In contrast to the no-NTA results of Fig. 2B, the presence of NTA in the medium kept Pu3+ in a soluble form in Fig. 4B.
While the desired end product is a bio-precipitated plutonium phase, modeling interpretation of experimental results shows that strong complexing ligands form soluble plutonium phases that are mobile, thus defeating an immobilization strategy. Such undesired reduced products were also observed upon reductions of uranium—forming U(IV)–citric acid complex (Francis and Dodge 2008)—and plutonium—forming Pu3+–EDTA (Boukhalfa et al. 2007) and Pu3+–NTA (Rusin et al. 1994).
We used biogeochemical modeling of experimental results in Reed et al. (2007) to advance our understanding of the fate of Pu in the absence and presence of Fe3+–NTA under anaerobic conditions. In all experiments, S. alga reduced PuO2+ to Pu3+, and evidence indicates that the Pu3+ could be precipitated as PuPO4(am). Modeling simulations support that reduction in the absence of Fe3+–NTA was from direct biotic PuO2+ reduction, but that PuO2+ reduction in the presence of Fe3+–NTA was due to an abiotic reduction reaction by biogenically produced Fe2+. These results explain that PuO2+ reduction was slowed in the presence of Fe3+–NTA because the bacteria preferentially reduced Fe3+ to Fe2+, which then abiotically reduced Pu(V) stepwise to Pu(IV) and then to Pu(III). Modeling results also show that the degree of PuPO4(am) precipitation depended on the concentration of the strong complexing ligand NTA. While precipitation out-competed complexation when NTA had an equimolar or smaller concentration compared to Pu, an excess of NTA could completely prevent precipitation of PuPO4(am).
The authors are grateful to Los Alamos National Laboratory for laboratory facilities. The research was supported, in part, by Environmental Remediation Sciences Program (ERSP) of the United States Department of Energy.