Interaction of uranium (VI) with bacteria: potential

Interaction of uranium (VI) with bacteria: potential applications in bioremediation of U contaminated oxic environments Sangeeta Choudhary & Pinaki Sar

Reviews in Environmental Science and Bio/Technology ISSN 1569-1705 Rev Environ Sci Biotechnol DOI 10.1007/s11157-015-9366-6

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Author's personal copy Rev Environ Sci Biotechnol DOI 10.1007/s11157-015-9366-6

MINI REVIEW

Interaction of uranium (VI) with bacteria: potential applications in bioremediation of U contaminated oxic environments Sangeeta Choudhary . Pinaki Sar

Ó Springer Science+Business Media Dordrecht 2015

Abstract Uranium and other metallic wastes released due to geochemical and several anthropogenic activities cause enormous damage to the environment. The fate and mobility of uranium (U) in the environment is affected by diverse microorganisms which interact through different mechanisms. Uranium at contaminated sites exists predominantly in two most common and stable valence states forms—the most oxidized valence state U(VI) exists as the highly soluble and toxic uranyl species (UO22?) while the reduced insoluble and less mobile, U(IV) is stable in the form of the mineral uraninite (UO2) under anoxic conditions. Reduced U(IV) species is less toxic and poorly soluble but it is liable to reoxidation and subsequent remobilization to soluble and more toxic U(VI) under oxic conditions. Fundamental understanding of nonreductive bacterial interaction mechanisms affecting the mobility and solubility of U(VI) in the environment is useful for developing suitable remediation and long-term management plan for U-contaminated sites. The present study gives an overview of various nonreductive bacterial interaction

S. Choudhary (&) Department of Bioscience and Biotechnology, Banasthali University, Rajasthan 304022, India e-mail: [email protected] P. Sar Department of Biotechnology, Indian Institute of Technology, Kharagpur 721302, India

processes which affects the mobility and solubility of U(VI) in oxygenic environments. Keywords Uranium Biomineralization Bioremediation Bacteria

1 Introduction Uranium (U), the heaviest (Z = 92) chemical element found in the nature, generally exists as triuranium octaoxide (U3O8) in mineral pitch blende. Natural U exists in three isotopic forms (U234, U235 and U238), the two most abundant isotopes, U235 (0.72 %) and U238 (99.27 %) have radioactive half-lives of about 7 9 108 and 4.4 9 109 years, respectively (Todorov and Ilieva 2006). Uranium is naturally present in the Earth’s crust, and is continuously released into the environment from various geochemical activities including weathering of rocks and minerals. In addition to such natural release, anthropogenic sources such as U-mining and milling operations including nuclear fuel cycle, uranium conversion, fuel fabrication, use of phosphate fertilizers and other industrial applications generate substantial quantities of U containing wastes (Lloyd and Macaskie 2000; Meinrath et al. 2003; Gavrilescu et al. 2009). Improper disposal and poor management of large amount of heavy metal and radioactive containing mixed wastes released from U mining and milling and other activities cause enormous damage to the environment

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by affecting surface and groundwaters, soils, subsurface sediments and even the catchment areas of drinking water (Gavrilescu et al. 2009). The mobility of U varies within the multiple subsurface zones that contain residual contaminant. Principal subsurface zones below the land surface include (1) the vadose zone, (2) capillary fringe zone through which the water table rises and falls, (3) the saturation zone containing groundwater. Along with the flow of groundwater, U contaminants mix with river water beneath the shoreline and finally get transferred to several impact receptors (Fig. 1). Major controlling factors affecting U mobilization includes the form of the residual U (such as mineral crystals, amorphous precipitates/coatings on the sediment), the transporting medium which can be infiltration of water from the land surface, flow of groundwater, and the rate of exchange between the form and transporting medium (Brown et al. 2008). Uranium at contaminated sites exists predominantly in two most common and stable valence states forms—U(VI), the most oxidized valence state, primarily as the most soluble and toxic uranyl species (UO22?) and the reduced insoluble and less mobile

U(IV), which is stable in the form of the mineral uraninite (UO2) under anoxic conditions. Recently different non-uraninite U(IV) phases {ningyoite [CaU(PO4)2], U2O(PO4)2, and U2(PO4)(P3O10)} have also been reported (Bargar et al. 2013). However, mineral uraninite is highly susceptible to oxidation under oxidizing conditions to highly soluble and mobile U(VI) form. The isotopes of U (238, 234, 235) mainly emit alpha particles that have little penetrating ability, the main hazard occurs when U compounds are ingested or inhaled. In comparison to tetravalent U, relatively soluble hexavalent U compounds are more likely to be a systemic toxicant (Gavrilescu et al. 2009). Several physical/chemical methods used to treat U-contaminated sites are expensive; an alternative to these technologies is the use of native subsurface bacteria for immobilizing U in contaminated groundwater and soil (Merroun and SelenskaPobell 2008). The speciation of uranium in aqueous systems under several environmental conditions is one of the critical factors affecting uranium bioremediation. Uranium (VI) aqueous speciation varies depending on the pH, ionic strength, U(VI) concentration and

Fig. 1 Schematic diagram showing U contamination path and mobility from disposal site to land surface, Vadose zone (unsaturated zone), phreatic zone (saturated zone) reaching

groundwater and again passing to water body (river, pond) along with flow of groundwater from where it reaches to several impact receptors (living organisms) on land and water

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presence of carbonate ions. Several speciation studies have reported that at pH values less than 3, U(VI) is present exclusively in the form of the uranyl cation, UO22?. In carbonate free systems an increase in solution pH favors the formation of positively charged hydroxo-uranium (VI) complexes that are transformed to negatively charged ones with a further increase in pH. However, carbonato-uranyl species are more favourably formed in highly alkaline conditions containing high concentrations of carbonaceous complexing agents than the hydroxide ions (Bargar et al. 1999; Gorman-Lewis et al. 2005). The tendency for these uranyl complexes to remain in solution has led to the common assumption that such aqueous complexation causes a high degree of uranium mobility. Furthermore, the mobility of these uranyl complexes is one major challenge to various proposed remediation strategies which rely on the binding of these complexes onto mineral surfaces and negatively charged bacterial surfaces. The fate of metallic contaminants released in the environment is controlled by diverse microorganisms which represent a significant fraction of the reactive surface area for interaction with dissolved molecules/nano minerals (Francis et al. 2004; Merroun et al. 2006). Four basic mechanisms by which bacteria can immobilize U are (1) microbial mediated reductive precipitation of U(VI) to U(IV), (2) U uptake and accumulation by cells, (3) biosorption and complexation with proteins, polysaccharides and microbial biomolecules, and (4) biomineralization of U(VI) with phosphates, carbonates (Fig. 2). All such interaction mechanisms that eventually alter the toxicity and mobility of U are subject of intense research for their applications in bioremediation. Uranium bioremediation approaches relying on microbial reduction of U(VI)–U(IV) have been studied extensively (Lovley et al. 1991; Shelobolina et al. 2004; Orellana et al. 2013; Converse et al. 2013). However, in oxic environments that do not favor reductive precipitation it may result in only temporary results if the redox conditions cannot be sustained. Therefore, resistance of biologically reduced U(IV) to reoxidation and subsequent remobilization to soluble U(VI) species is significant for long term bioremediation results. Previous reports suggest that various conditions such as presence of carbonate, nitrate, Fe(III) minerals, manganese oxides or organic ligands such as citrate and EDTA may favor U(IV) reoxidation even in

anoxic conditions (Luo and Gu, 2011; Singh et al. 2014). Since environmental bacteria can interact with U to modify its speciation and mobility, U in turn exerts a permanent pressure which may influence the structure and metabolic activity of bacterial communities; therefore fundamental understanding of these interactions will be useful for developing appropriate remediation and long-term management plan for U-contaminated sites. In recent years considerable efforts have been made to explore the microbial community structure and composition using both culture -dependent and—independent approaches at various U-mine contaminated habitats. Various bacterial strains have been obtained from sediment and groundwater samples collected at the Oak Ridge Field Research Center, as well as from several milling and mining impacted environments (North et al. 2004; Cardenas et al. 2008; Dhal and Sar 2014; Kumar et al. 2013; Islam et al. 2014). In number of cases bacterial strains isolated from such sites have been well characterized in terms of their interaction mechanisms leading to U biosorption/bioaccumulation or precipitation (Selenska-Pobell et al. 1999; Panak et al. 2002; Merroun and Selenska-Pobell 2008; Choudhary and Sar 2011; Newsome et al. 2014). In comparison to reductive precipitation of U(IV), interaction of U(VI) with surface and subsurface aerobic microorganisms is relatively less attended that may play an useful role to control U toxicity in low pH, high nitrate and aerobic conditions. However, our knowledge on uranium biogeochemistry is still incomplete to develop efficient predictive models to study long-term behavior of U in such contaminated environments (Hazen and Tabak, 2005). This review presents how the mobility and toxicity of U(VI) in the environment is affected by nonreductive bacterial interaction processes which may have potential to bioremediate contaminated sites. 1.1 Biosorption and intracellular accumulation (bioaccumulation) leading to U- complexation (with metal-binding peptides, proteins, polysaccharides and other biomolecules) Biosorption of metals by bacteria is one of the most frequent and important events that influences the soluble metal concentration in the environment. Biosorption is the metabolism- independent passive accumulation of metals and radionuclides by

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Fig. 2 Schematic diagram showing the mechanisms of bacteria-U interactions [U(VI) reduction, biosorption and complexation, biomineralization, uptake and accumulation] and subsequent removal. OM (outer membrane), CM (cytoplasmic membrane), PS (periplasmic space), Gly-P (glycerol phosphate), UO22? (uranyl ion U(VI) form), M phos (Microbial phosphatases pH 4.0-7.0). a Reoxidation (in presence of nitrate, Fe(III) hydroxides, low pH). b Biosorption (Interaction with various negatively charged ligands on cell surface).

and meta-autunite minerals).

Complexation with surface i.e.

biomolecules EPS (Extracellular Polysaccharide), S-layer protein.

Immobilization of U(VI) by poly phosphate granules in

cytoplasm. References (Lovley et al. 1991; Merroun and Selenska-Pobell 2008; Choudhary and Sar 2011)

Precipitation as U(VI) phosphate minerals (Autunite

microbial cells while bioaccumulation refers to intracellular accumulation of metals by alive cells and can be metabolism-dependent. Bioaccumulation occurs in two stages, in which an initial rapid metal binding is followed by a relatively slow intracellular accumulation achieved by active transport (Aksu and Do¨nmez 2000). Two types of uptake system are involved for heavy-metal ions: one is fast, unspecific, chemiosmotic gradient-dependent and remains constitutively expressed. The second type of uptake system is substrate-specific, slow process and often uses ATP hydrolysis as the energy source (Nies 1999). In living cells, apart from having specific or nonspecific metal uptake channels/transporters, metabolic activity may also influence biosorption/bioaccumulation by changing pH, organic and inorganic constituents (Gadd 2004). Metal accumulation by biosorption or

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bioaccumulation may also accompany or lead to nucleation, precipitation and biomineral formation. However unlike metabolically essential metals, U accumulation mostly occurs via metabolism-independent processes mostly due to increased membrane permeability as a result of its toxicity (Suzuki and Banfield 1999). Various studies have been performed to quantify U binding among bacteria from diverse origin including laboratory strains (Panak et al. 2002; Francis et al. 2004; Nakajima and Tsuruta 2004; Kazy et al. 2009) as well as environmental isolates (Merroun and Selenska-Pobell 2008; Llorens et al. 2012; Lu¨tke et al. 2012). These studies demonstrated that U biosorption is dependent on several physico-chemical properties as well as on the specific microbial species. In order to understand the mode and mechanism of


metal ion binding by the bacterial cell, several macroscopic and molecular-level studies have been conducted. Surface complexation model used to quantify the site densities, deprotonation constants, and metal-binding constants of the functional groups present on bacterial surfaces demonstrated that the carboxylic acid and the phosphate groups are the major binding ligands involved in the coordination of the radionuclide (Fein et al. 1997; Fowle et al. 2000; Kelly et al. 2002; Lu¨tke et al. 2012; Gorman-Lewis et al. 2005). Role of carboxyl and phosphate groups present in bacterial S-layers proteins (the outermost cell envelope component of many bacteria) in U binding has also been demonstrated in Lysinibacillus sphaericus JG-A12, JG-B53 isolated from a U-contaminated environment (Merroun and Selenska-Pobell 2008; Lederer et al. 2013). Various techniques, which have been employed for identification of binding sites and ligands include several spectroscopic and microscopic techniques such as Fourier Transform Infrared (FTIR) spectroscopy, X-ray Absorption Spectroscopy (XAS), specially Extended X-ray Absorption Fine Structure (EXAFS) spectroscopy, X-ray Photoelectron Spectroscopy (XPS), Time-Resolved Laser-induced Fluorescence Spectroscopy (TRLFS), Energy Dispersive X-ray (EDX) analysis, Scanning Electron Microscopy (SEM), Transmission Electron Microscopy (TEM), X-Ray Diffraction (XRD) analysis (Andres et al. 1994, Chojnacka 2010; Merroun and Selenska-Pobell 2008; Templeton and Knowles 2009; Barkleit et al. 2011; Choudhary and Sar 2011). Bacterial cells have developed several mechanisms to immobilize and precipitate U and other metals once they are accumulated intracellularly. A well studied process is U chelation by polyphosphate bodies which are linear polymers of inorganic phosphate (Pi) residues linked by phosphoanhydride bonds, with chain length varying from 3 to 1000 Pi residues (Van Veen et al. 1993). It has been suggested that the intracellular chelation of heavy metals by polyphosphate is a passive heavy metal tolerance mechanism (Keasling and Hupf 1996). Uranium chelation by polyphosphate bodies has been observed in bacterial strains such as Acidithiobacillus ferroxidans and Sphingomonas sp. S15-S1 isolated from different U contaminated sites (Merroun et al. 2003, 2006). In a number of studies U(VI) removal capacities of poly-P producing Pseudomonas aeruginosa or mixed bacterial consortium were evaluated for their application in

simultaneous phosphate removal and U(VI) precipitation (Renninger et al. 2004). Microorganisms produce a range of specific and nonspecific metal-binding compounds. Nonspecific metal-binding compounds range from simple organic acids and alcohols to macromolecules, such as polysaccharides, humic and fulvic acids (Birch and Bachofen 1990; Gadd 2004). Extracellular polymeric substances (EPS) produced by bacteria may provide potential sites for the sequestration of metal ions (Beech and Cheung 1995). EPS of bacterial strains isolated from different extreme (U mining wastes, U mill tailings, and groundwater of radioactive repositories) and non-extreme habitats showed high ability to accumulate U and other heavy metals (Kazy et al. 2002; Merroun and Selenska-Pobell 2008). Uranium phosphate biomineralization as a result of complexation of U within the cellular EPS has been studied in several strains (Pseudomonas stutzeri DSMZ 5190, A. ferroxidans, Myxococcus xanthus) (Merroun and Selenska-Pobell 2008). Siderophores are the chelating agents produced by bacterial cells for iron assimilation (Gadd 1992). Siderophores are also able to bind other metals such as Mg, Mn, Cr, Ga and radionuclides such as Pu (Birch and Bachofen 1990). Siderophore-mediated U(VI) sequestration by marine cyanobacterium Synechococcus elongatus BDU 130911 have been recently reported (Rashmi et al. 2013). Because of such metal-binding abilities, there are potential applications for siderophores in reprocessing of nuclear fuel, bioremediation of metal-contaminated sites and treatment of industrial wastes (Renshaw et al. 2002). 1.2 Biomineralization Microorganisms are able to precipitate metals and radionuclides under the form of carbonates, phosphates and hydroxides, as these ligands are concentrated near the cell surface, which provides nucleation foci for precipitation (Lloyd and Macaskie 2000). Enzymatically-generated phosphate residues precipitate a range of metals and radionuclides as phosphate compounds (Macaskie et al. 1994; Yong and Macaskie 1995). The membrane bound phosphatase enzyme can liberate inorganic phosphate from an externally supplied organic phosphate donor (e.g., Glycerol 2- phosphate) precipitating metals/radionuclides as phosphates on the biomass. Initially, this

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system was developed for bioaccumulation of Cd and Zn ions by a Citrobacter strain (Macaskie et al. 1992); later on it was used for precipitation of U and other actinides, lanthanides and heavy metals (Lloyd and Macaskie 2000). Phosphatase-mediated U precipitation as autunite/meta-autunite mineral phases [U(VI) phosphate compounds)] have been reported in various strains isolated from radionuclide and metal contaminated environments (Merroun et al. 2003, 2006; Koban et al. 2004; Martinez et al. 2007; Nedelkova et al. 2007; Choudhary and Sar 2011, Sousa et al. 2013). In natural U-containing soils and sediments, bacterial cells have been observed that were entirely covered with U phosphate minerals suggesting possible role of environmental bacteria in the formation of mineral phases of U in the soils by a biomineralization process which may slowly lead to bioattenuation (Mondani et al. 2013). The co-precipitation of U with carbonate minerals (calcite) has also been demonstrated using spectroscopic and luminescence methods (Reeder et al. 2001).

2 Application of U(VI) biomineralization for bioremediation Uranium biomineralization has been investigated for potential use at the various U contaminated sites (Beazley et al. 2007; Martinez et al. 2007, Choudhary and Sar 2011). Phosphatase mediated bioprecipitation of U was investigated with Bacillus and Rahnella strains in presence of Glycerol-3-phosphate. The U-phosphate precipitate formed by these isolates was identified as calcium autunite [Ca(UO2)2(PO4)2]. Further investigations revealed that the Rahnella strain could also biomineralise uranyl to chernikovite [H2(UO2)2(PO4)2] under anaerobic conditions and in the presence of high nitrate concentration (Beazley et al. 2009). Another study with bacterial isolates (Aeromonas hydrophila, Pantoea agglomerans and Pseudomonas rhodesiae) from circumneutral-pH groundwater [from Area 2 of the U.S. Department of Energy, Oak Ridge Field Research Center (ORFRC)] having very high concentration of dissolved calcium demonstrated U precipitation in both aerobic as well as nitrate reducing conditions via Glycerol-3-phosphate hydrolysis. TEM and EDX of U precipitates showed that U(VI) was preferentially incorporated into aggregates of

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nanocrystals of hydroxyapatite, Ca5(PO4)3OH (Shelobolina et al. 2009). Uranium biomineralization by a U-resistant P. aeruginosa strain J007 isolated from U mine waste site of India was characterized for its potential in bioremediation (Choudhary and Sar, 2011). Microcosm studies using mine effluent containing high amount of soluble U (3800 gl-1) was performed to evaluate the effectiveness of this bacterium in field application for U bioremediation. Bioaugmentation with P. aeruginosa J007 cells stimulated nearly 99 % removal of soluble U under in situ conditions. Further XRD data confirmed that the bacterium precipitated U as crystalline U phosphate compounds [UO2(PO3)2, (UO2)3(PO4)2H2O and U2O(PO4)2] compounds. Apart from studies of U biomineralization using pure culture isolates, the role of indigenous soil bacteria in Glycerol-3-phosphate-mediated U precipitation have been demonstrated at U contaminated soils from the Department of Energy (ORFRC) in small flow-through columns under aerobic conditions at pH 5.5 and 7.0. High phosphate concentration was detected in both the columns amended with Glycerol3-phosphate in comparison to control conditions indicating the stimulation of phosphatses producing bacteria in response to organic phosphate source. Solid-phase extraction scheme and XAS analysis confirmed U(VI) immobilization in both pH 5.5 and pH 7 columns through the formation of stable uranyl phosphate minerals (Beazley et al. 2011). Another microcosm-based study demonstrated the efficiency of biomineralization of U(VI)–phosphate minerals over bioreduction of U(VI) to U(IV) under anaerobic condition at pH 5.5 and 7.0. Simultaneously, wet chemical extractions and EXAFS analysis of these sediments indicated rapid removal of dissolved U(VI), and the simultaneous existence of U–P species in the solid phase was further confirmed that U was precipitated as U(VI)–phosphate minerals in sediments amended with Glycerol-2-phosphate but there was no evidence for U(VI) reduction (Salome et al. 2013).

3 Conclusion Extensive research has been done on U biogeochemistry and bioremediation, nevertheless considering the management and safety of future nuclear waste it is


very important to understand the mobility and interaction of U(VI) with diverse microorganisms having endogenous genetic, biochemical and physiological properties. Both cell surface sequestration as well as indirect interaction of U(VI) with phosphates, carbonates or different biomolecules which may lead to complexation/biomineralization have great potential to affect the U toxicity and mobility. Employment of U(VI) biomineralization has prospective application in surface and subsurface environments under oxic conditions where it is difficult to maintain the stability of bioreduced U(IV). Some of the studies using pure bacterial cultures have demonstrated removal of U(VI) via generation of poorly soluble uranyl phosphates. However, U-phosphate biomineralization application in in situ field studies is mostly reliant on use of Glycerol phosphate, therefore future studies in this direction are required to find cost-effective alternatives. In current time bacteria-U(VI) interaction process appears to be a promising approach for U bioremediation as various surface complexation, microscopic, spectroscopic and molecular techniques are highly advanced and can be used to observe and improve their application in the field. Considering public safety and continued use of uranium as a source of balanced energy production, there is requirement for manifold management approaches for U contaminated wastes sites such as biostimulation, bioaugmentation, and other bioremediation approaches.

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