Keywords

1 Introduction

The release of radioactive materials into the environment as a consequence of ongoing nuclear activities and through accidents (most recently Fukushima) has led to the accumulation of radioactive waste with tremendous environmental impact (Kazy et al. 2009). The major burden of anthropogenic radioactivity are, however, actinides (U, Pu, and Np) released from nuclear fuel processing plants and the fission products I, Cs, Sr, and Tc (Geissler et al. 2010). The radionuclides released may be in various chemical forms and may eventually transform in the environment to oxide, coprecipitates, and ionic, inorganic, and organic complexes (Francis and Dodge 2009).

Many remediation technologies exist to treat these contaminants to reduce risk to humans and the environment. The removal of radionuclides is mostly achieved by their reduction using physiochemical methods, such as the pump method, permeable reactive barriers (PRBs), ligand coupling, selective ion exchange separation, and complex formation with organic acids. However, environmental friendly methods are in demand due to high cost, low sustainability, and potential hazards linked with chemicals (Mohapatra et al. 2010). A lot of efforts have been made to understand the bioremediation of radionuclides as a potential remediation strategy. The main emphasis is to develop lower-cost techniques with minimal environmental disruption for removal of radionuclide contaminants.

Bioremediation technology uses microorganisms (prokaryotes and eukaryotes) to reduce, eliminate, or contain contaminants present in the environment (McCullough et al. 1999). The use of microbes for the remediation is determined largely by their ability to survive and catalyze the desired function under high radiation stress (Brim et al. 2000). The remediation strategy may be adopted by using techniques designed to extract or segregate the contaminated fraction from the rest of the soil, either in situ or ex situ. Ex situ bioremediation involves excavating the contaminated material and its treatment in aboveground facilities located on-site or off-site, whereas in situ remediation is undertaken at the site of contamination. In situ remediation technologies have several advantages over ex situ methods as they are cheaper and less disruptive because no excavation is required; however, they may require longer treatment time frames than the ex situ techniques (Francis and Nancharaiah 2015). Microbial activity could change the concentrations of radionuclides and their bioavailability in environment by changing chemical nature such as speciation, solubility, and sorption properties (Francis 2007). Microorganisms are ubiquitous in soil, sediments, and subsurface environment although their population sizes and respiratory metabolic activities can vary considerably depending on energy and nutrient inputs. They have developed very effective protection mechanism against radionuclides based on their ability for enzymatic/reductive precipitation, solubilization, biosorption, bioaccumulation, etc. (Merroun and Pobell 2008). Theoretically both immobilization and mobilization present opportunities for bioremediation. Although mobilization can be a better long-term strategy, the risk involved to immobilize the toxic contaminant in the capture zone makes it unsuitable to deploy in field studies (Cao et al. 2010). Among the several microbial processes that determine the environmental fate of radionuclides, biotransformation has been emerging as a potential technique achieved by either direct enzymatic reduction or by their end products (Mohapatra et al. 2010).

This chapter documented the existing fundamental information about critical redox reactions that lead to immobilization, understanding of key microbial metabolic processes during biotransformation, and the evaluation of results on microbial metabolic processes. Further aspects and new developments in genetic engineering are examined.

1.1 Uranium

Uranium exists in several oxidation states. U(IV) and U(VI) are the predominant states found in the environment (Francis 2002). Uranium exhibits higher solubility in oxidized form as uranyl ion (UO2 2+). When uranyl ions encounter a reducing agent such as organic matter and microorganisms, it is precipitated into considerably less soluble lower oxidation state as the oxide mineral uraninite (UO2) (McCullough et al. 1999; Marshall et al. 2010). Both aerobic and anaerobic microorganisms have been reported to involve directly or indirectly in redox transformation of various chemical forms of uranium in the environment (Francis 2012):

$$ \mathrm{U}\left(\mathrm{V}\mathrm{I}\right)\ \overset{\mathrm{dissimilatory}\ \mathrm{metal}\ \mathrm{reducers},\ \mathrm{fermenters},\ \mathrm{sulfate}\ \mathrm{reducers}}{\to}\mathrm{U}\left(\mathrm{I}\mathrm{V}\right) $$

Immobilization of U has been potentially achieved by a diverse range of bacteria and archaea; however, the most identified are dissimilatory Fe(III) or sulfate-reducing bacteria (DIRB and DSRB), γ and δ Proteobacteria, and Firmicutes in the presence of suitable electron donors leading to bioreduction and bioprecipitation (Loughlin et al. 2011). At low pH, uranium adsorption onto cell surface of Bacillus subtilis (Lloyd et al. 2002) and cell wall of fungus Rhizopus arrhizus involving coordination of amine N of chitin (Gadd 2002) has been reported. The precipitation of uranium by Citrobacter sp. has also been demonstrated with enzymatically liberated inorganic phosphate (Macaskie et al. 1992). Thus, uranyl phosphate accumulates as polycrystalline HUO2PO4, at the cell’s surface. Recently, Rui et al. (2013) have reported reduction of this biogenic mineral to U(IV) by metal-reducing bacteria, namely, Geobacter sulfurreducens strain PCA, Anaeromyxobacter dehalogenans strain K, and Shewanella putrefaciens strain CN-32.

Similar to Citrobacter, in Halomonas sp., U(VI) was not only bound to the cell surface but also accumulated intracellularly as electron-dense granules. EXAFS analysis of halophilic and non-halophilic bacterial cells demonstrated that U(VI) existed predominantly as uranyl hydrogen phosphate and complex forms of phosphate such as hydroxophosphate or polyphosphate as well as other ligands such as carboxyl species (Francis et al. 2004).

Direct enzymatic reduction depending on the hydrogen or organic compounds like acetate, ethanol, lactate, etc. on the surface of Shewanella putrefaciens, Desulfovibrio vulgaris, and Desulfovibrio desulfuricans in the presence of c-type cytochrome was reported as

$$ \begin{array}{l}\mathrm{U}{\mathrm{O}}_2{{}^{2+}}_{\left(\mathrm{soluble}\right)} + {\mathrm{H}}_2\to \mathrm{U}{{\mathrm{O}}_2}_{\left(\mathrm{insoluble}\right)}+2{\mathrm{H}}^{+}\\ {}2\mathrm{U}{\mathrm{O}}_2{{}^{2+}}_{\left(\mathrm{soluble}\right)} + \mathrm{C}{\mathrm{H}}_2\mathrm{O} + 2{\mathrm{H}}_2\mathrm{O}\to 2\mathrm{U}{\mathrm{O}}_{2\left(\mathrm{insoluble}\right)} + \mathrm{C}{\mathrm{O}}_2 + 2{\mathrm{H}}^{+}\end{array} $$

However, effective electron donor is selected on site-specific basis, e.g., ethanol was recommended for the Oak Ridge site, while acetate for US DOE Shiprock and Old Rifle, UMTRA site, Colorado (Finneran et al. 2002; Hazen and Tabak 2005; Luo et al. 2007). These electron donors stimulate nitrate reduction that can cause a rise in pH hence promoting removal of uranium either via hydrolysis and precipitation (Shelobolina et al. 2007) or bioreduction where uranium is being utilized as electron acceptor (Istok et al. 2004).

Indirect reduction by Fe(III)-reducing bacterium Geobacter sulfurreducens in the presence of homologous cytochrome (PpcA) and a triheme periplasmic cytochrome c7, anaerobic spore-forming fermentative bacteria Clostridium sp., mesophilic sulfate-reducing bacteria, hyperthermophilic archaea, thermophilic bacteria, and acid-tolerant bacteria has also been reported (Merroun and Pobell 2008; Francis and Dodge 2009; Anderson and Lovely 2002; Prakash et al. 2013).

1.2 Technetium

Technetium can exist in multiple oxidation states (0, +3, +4, +5, +6, +7). Tc(VII) has attracted considerable recent interest due to its mobility as soluble pertechnetate ion TcO4 − under oxic conditions and weak ligand-complexing capabilities like Np(V) (Lloyd et al. 2005; Francis 2007). Under reducing conditions, it can form insoluble and strongly sorbing hydrous phase Tc(IV)O2 (Newsome et al. 2014).

Chemolithotrophic, haloalkaliphilic, aerobic, facultative, and anaerobic bacteria have been identified to reduce Tc(VII) catalyzed by the hydrogen component of the formate hydrogenlyase complex (FHL) or c-type cytochrome. The dissimilatory reduction of Tc is favored in neutral, acidic, and alkaline environment depending on the type of microbial activity (Mohapatra et al. 2010) coupled to oxidation of H2 or certain organic compounds as an electron donor.

$$ \mathrm{T}\mathrm{c}{{\mathrm{O}}^{4-}}_{\left(\mathrm{soluble}\right)} + 3{\mathrm{e}}^{-} + 4{\mathrm{H}}^{+}\to \mathrm{T}\mathrm{c}{\mathrm{O}}_2.n{\mathrm{H}}_2{\mathrm{O}}_{\left(\mathrm{insoluble}\right)} + \left(2-n\right){\mathrm{H}}_2\mathrm{O} $$

A Halomonas strain isolated from a seawater removed Tc(VII) by aerobically reducing it to Tc(IV). The predominate reduced species include TcO2, Tc(OH), and TcS2. The reduction of Tc(VII) by indirect microbial processes such as biomineralization via biogenic sulfide and Fe(II) has also proved to be efficient and useful mechanism for remediation by Geobacter, Anaeromyxobacter, and Shewanella (Plymale et al. 2011). However, limited ingrowth of Fe(II) at high concentration of nitrate in Oak Ridge, Tennessee, USA, has been explored suggesting strong inhibition of metal reduction (Geissler et al. 2010).

1.3 Plutonium

Plutonium has complex redox chemistry. At neutral pH, Pu can exist in multiple oxidation states, but small changes in pH and redox conditions can lead to changes in speciation. Plutonium can coexist as Pu(IV), Pu(V), and Pu(VI) in oxic environment. With a slight increase in microbial activity, Pu (VI) and Pu(V) can be reduced to Pu(IV) due to very small difference in reduction potential as shown below (Francis 2007). Therefore, the Pu(IV) is the dominant oxidation state under most environments and forms highly insoluble Pu(OH)4.

figure a

Fe(III)-reducing bacteria like Shewanella oneidensis, S. putrefaciens, and Geobacter metallireducens have been reported for the enzymatic reduction of Pu (VI)/(V) to Pu(IV), while Pu(III) has very little production. But the presence of complexing ligands has been found to facilitate Pu(IV) to Pu(III) reduction. However, anaerobic bacteria Bacillus sp., Bacillus mycoides, Serratia marcescens (Luksiene et al. 2012), and Clostridium sp. have shown tendency for reductive dissolution of Pu(IV)–Pu(III) (Francis 2007; Francis and Dodge 2009). Shewanella alga can reduce Pu(V)O2 + to amorphous Pu(III)PO4 (Reed et al. 2007; Deo et al. 2011). Lichen Parmotrema tinctorum has been observed to play an important role in the elimination of Pu(VI) and U(VI). The accumulation of Pu and U(VI) has been identified on the upper and lower surfaces and in cortical and medullary layers, respectively. UV/vis absorption spectroscopy showing microbial activity can further reduce Pu(VI) to Pu(V) in solution and to Pu(IV) as Pu(IV) hydroxide by releasing organic substances, while the oxidation state of U(VI) remains unchanged (Ohnuki et al. 2004).

1.4 Neptunium

Neptunium predominantly exists as the neptunyl cation Np(V)O2 + in aerobic environment that has low ligand-complexing ability. It is highly mobile, is non-sorptive, and can be removed via combined bioreduction–biomineralization under anaerobic conditions. In situ bioremediation of Np(V) supplied with hydrogen or pyruvate is shown as (Rittmann et al. 2002; Mohapatra et al. 2010)

$$ \mathrm{N}\mathrm{p}{{\mathrm{O}}^{\mathrm{2}+}}_{\left(\mathrm{soluble}\right)} + {\mathrm{e}}^{-} + \mathrm{4}{\mathrm{H}}^{+}\to \mathrm{N}{{\mathrm{p}}^{\mathrm{4}+}}_{\left(\mathrm{insoluble}\right)} + \mathrm{2}{\mathrm{H}}_{\mathrm{2}}\mathrm{O} $$

Anaerobic microbial reduction of aqueous Np(V) citrate by G. metallireducens and S. oneidensis and reduction of un-chelated Np(V) by S. oneidensis, by S. putrefaciens MR-1, and by sulfate-reducing bacteria Desulfovibrio desulfuricans have been reported (Icopini et al. 2007). Contrary to reduction of UO2 2+ by first transferring an electron to an actinyl ion and then generation of U(IV) by disproportionation, it has been investigated that G. sulfurreducens cannot transfer an electron to NpO2 + (Renshaw et al. 2007).

The reduced Np(IV) can be precipitated with phosphate liberated by phosphatase activity of Serratia and Citrobacter sp. (Lloyd et al. 2000). Abiotic reduction has also been suggested in microbial active sediment system when provided with electron donor acetate or lactate (Newsome et al. 2014).

1.5 Strontium and Cesium

Since strontium and cesium are both redox inactive, their removal from the environment is mostly influenced by indirect microbial interactions such as bioaccumulation, sorption, and precipitation. Many active metabolizing microorganisms including Pseudomonas fluorescens, epilithic cyanobacteria, Halomonas, and Sporosarcina pasteurii have demonstrated ability to remove strontium as crystalline SrCO3 (Newsome et al. 2014). In most cases, this occurs through microbially induced calcium carbonate precipitation (MICP) (Fujita et al. 2010) that can be implemented using reduction of sulfate, bacterial ureolysis, fermentation of fatty acids, and denitrification process (Francis and Nancharaiah 2015).

  • Microbially catalyzed hydrolysis of urea produces ammonia and bicarbonate and increases the pH. Bicarbonate promotes the precipitation of CaCO3, whereas ammonium ions promote the exchange of sorbed Sr (Fujita et al. 2000, 2004).

    $$ \begin{array}{l}\mathrm{N}{\mathrm{H}}_2\left(\mathrm{C}\mathrm{O}\right)\mathrm{N}{\mathrm{H}}_2 + 3{\mathrm{H}}_2\mathrm{O}\overset{\mathrm{Urease}}{\to }2\mathrm{N}{{\mathrm{H}}_4}^{+} + \mathrm{H}\mathrm{C}{{\mathrm{O}}_3}^{-} + \mathrm{O}\mathrm{H}\\ {}\mathrm{S}\mathrm{olid}\ \hbox{-}\ \mathrm{S}{\mathrm{r}}^{2+} + 2\mathrm{N}{{\mathrm{H}}_4}^{+}\leftrightarrow \mathrm{S}\mathrm{olid}\ \hbox{-}\ \left(\mathrm{N}{{\mathrm{H}}_4}^{+}\right) + \mathrm{S}{\mathrm{r}}^{2+}\end{array} $$

  • Denitrification involves the reduction of NO3 − to N2 in the presence of an electron donor (i.e., acetate) (Wu et al. 2011).

    $$ \begin{array}{l}8/5\ \mathrm{N}{{\mathrm{O}}_3}^{-} + \mathrm{C}{\mathrm{H}}_3\mathrm{C}\mathrm{O}{\mathrm{O}}^{-}\overset{\mathrm{Denitrifiers}}{\to }4/5\ {\mathrm{N}}_2 + 2\ \mathrm{C}{\mathrm{O}}_2 + 13/5\ \mathrm{O}{\mathrm{H}}^{-} + 1/5\ {\mathrm{H}}_2\mathrm{O}\\ {}\mathrm{C}{\mathrm{O}}_2 + {\mathrm{H}}_2\mathrm{O}\leftrightarrow \mathrm{H}\mathrm{C}{{\mathrm{O}}_3}^{-} + {\mathrm{H}}^{+}\\ {}\mathrm{S}{\mathrm{r}}^{2+} + \mathrm{H}\mathrm{C}{{\mathrm{O}}_3}^{-} + \mathrm{O}{\mathrm{H}}^{-}\leftrightarrow \mathrm{S}\mathrm{r}\mathrm{C}{\mathrm{O}}_3\left(\mathrm{s}\right) + {\mathrm{H}}_2\mathrm{O}\end{array} $$

  • Fermentation of fatty acids like citric acid (Anderson and Appanna 1994) due to the production of CO2 from citrate metabolism.

In all abovementioned reactions, strontium is integrated into the calcite structure by substituting for calcium, thus forming strontium carbonate minerals of very low solubility (Fujita et al. 2008).

Alternative techniques proposed for strontium include sequestration as (Ba,Sr)SO4 by desmid green algae (Krejci et al. 2011), incorporation into biogenic hydroxyapatite produced by a Serratia sp. (Handley-Sidhu et al. 2011a, b), and co-treatment with Tc(VII) and high ammonium and nitrate levels via microbially induced increased in pH and alkalinity and during bioreduction of Fe(III) and nitrate (Thorpe et al. 2012).

The mechanism of cesium removal by cesium-accumulating bacteria (strains Rhodococcus erythropolis CS98 and Rhodococcus sp. strain CS402) isolated from soil cannot be modeled as simple adsorption as they display the rod–coccus growth cycle and contain meso-diaminopimelic acid, mycolic acids, and tuberculostearic acids (Tomioka et al. 1992). Since the same metabolic transport systems are involved in K+ and Cs+ absorption due to the chemical similarity of the cations, the presence of potassium inhibited Cs accumulation by these strains (Tomioka et al. 1994).

Recently, adsorption of cesium by unicellular green algae and heterocystous blue green algae has been reported (Shimura et al. 2012). The terrestrial cyanobacterium Nostoc (N) was found to form jelly-like clumps of polysaccharides that absorbed high levels of radioactive cesium (Sasaki et al. 2013). It was monitored that in Nihonmatsu City, Fukushima Prefecture, N. commune accumulated 415,000 of 134Cs and 607,000 Bq kg−1 of 137Cs dry weight, respectively (Fujita et al. 2013). Very high radiocesium activities have been observed in the fruiting bodies of several fungal species since the Chernobyl accident (Dighton and Horrill 1998). The bonding of Cs with the cells (extra- or intracellular) remains unclear, as does the long-term fate of bio-associated Cs. However, in sediment systems, the Cs is fixed at interlayer sites and frayed edge sites of phyllosilicate minerals where Cs can form inner sphere complexes. Indirect microbial impacts such as the release of competing ammonium ions or changes in mineral stability may play a role in controlling Cs mobility (Brookshaw et al. 2012).

1.6 Iodine

Microbial processes can significantly influence the mobility of iodine in the environment through volatilization in the form of organic iodine compounds (i.e., CH3I), the oxidation of iodide (I–) to iodine (I2), the bioreduction of iodate (IO3 −) to iodide (I−), and via bioaccumulation (Amachi 2008; Jabbar et al. 2013).

Pure cultures of bacteria and fungi have shown tendency to incorporate radioiodine in the presence of organic substances or biomass via irreversible sorption process (Bors and Martens 1992). Positively charged single sites were recognized to be responsible for iodide sorption on to cell wall of Gram-positive soil bacterium Bacillus subtilis. However, the uptake and accumulation of iodide in washed cell suspensions of marine bacteria was accelerated by the addition of glucose without exhibiting any effect on iodate deposition (Amachi et al. 2005). Recently, the adsorption of iodine by unicellular green algae has been reported (Shimura et al. 2012).

D. desulfuricans was able to enzymatically reduce iodate to iodide in bicarbonate and HEPES buffers, while S. putrefaciens was only able to perform this transformation in HEPES (Councell et al. 1997). As Fe(II), sulfide, and FeS were shown to abiotically reduce iodate to iodide, it is inferred that Fe(III)- and sulfate-reducing bacteria are able to mediate iodate reduction both directly and indirectly (Hu et al. 2005). This study was further illustrated by sorption of 54 % of iodine and 75 % of 129I from waters at near-neutral pH in in situ bacteriogenic Fe(III) oxides at Chalk River, Canada (Kennedy et al. 2011).

2 Mechanisms of Bioimmobilization of Radionuclides

The role of microbes in determining the environmental fate of radionuclides and understanding microbe–radionuclide interactions is of great importance for developing effective methods of bioremediation. In this section we discussed the principles and key factors affecting the processes of microbial transformations of radionuclides. The immobilization is achieved by direct or indirect reduction, biosorption onto the cells, precipitation by organic complexing ligands released by the cells, and bioaccumulation (Fig. 1) (Cao et al. 2010) depending on the form of the metal, the availability of electron donors and acceptors, nutrients, and environmental factors including moisture, temperature, pH, and Eh (Francis and Nancharaiah 2015).

Fig. 1
figure 1

Mechanisms of microbe–uranium interactions

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2.1 Biosorption

Biosorption describes the passive uptake of radionuclide species to the surface of living or dead microbial cells by metabolism independent of physicochemical mechanisms. Metal cations either adsorb on cell wall possessing electronegative charges or may bind through chemical sorption by ligands such as carboxyl, amine, hydroxyl, phosphate, etc. The process is species specific and is affected by the chemical speciation of metal, pH of medium, physiological state of the cells, as well as the presence of exopolymers (Merroun and Pobell 2008). Biosorption is perhaps best suited to remove radionuclides of low concentration from effluents because it is faster than bioaccumulation and the biosorbent is easy to regenerate.

However, limitations included desorption due to competition with other cations for binding sites and saturation preventing further sorption leading to an inadequate long-term solution for bioremediation.

2.2 Bioaccumulation

Bacteria, fungi, and algae have been identified to have ability to accumulate radionuclides extra- or intracellularly independent of metabolism due to similarities with some essential elements needed for cell functioning. The accumulation mechanism may involve either chelation by polyphosphate granules or formation of needle-like crystals supported by enhanced cell membrane permeability resulting from the toxic effects of radionuclides (Cao et al. 2010; Francis and Nancharaiah 2015). However, metabolism-dependent transport process via siderophore-mediated Fe(III) transport system has been reported.

2.3 Biomineralization/Bioprecipitation

Radionuclide precipitation through microbial generation of ligands and biogenic mineral formation play an important role in retention of contaminant. The enzymatically produced highly localized ligand concentration around the cells provides the nucleation site for precipitation, e.g., phosphate precipitation. Contrary to direct precipitation, biogenic mineral method involves precipitation of a metal biomineral by microbial cells which then act as a trap for the subsequent deposition of radionuclide of interest.

2.4 Bioreduction

Microbes offer great potential for controlling mobility of target radionuclides in contaminated environment through reduction of radionuclides that are redox active and less soluble when reduced. Since microbial life is based on chemical energy driven by metabolic process that requires a reduced electron donor and an oxidized electron acceptor (Pedersen 2005), the reduction can be carried out directly by microbes or indirectly through electron transfer by dissimilatory metal-reducing bacteria (DMRB). Different mechanisms may exist for electron transfer from donor to acceptor among different species. It may or may not involve direct contact. Although U(VI) is significantly soluble to have direct contact with a cell, bacteria must have mechanisms of transferring electron extracellular to acceptor Fe(III) or sorbed U(VI). These include the use of electron-carrier protein such as cytochrome exposed to the outside of the cytoplasmic membrane, within the periplasm, and/or in the outer membrane or through release of chelators/electron shuttle like flavins and through conductive cell surface appendages such as pili (nanowires) (Lovley et al. 2004; Nielsen et al. 2010; Pfeffer et al. 2012).

Some DMRB can be beneficial by conserving energy for anaerobic growth via radionuclide reduction as in case of Fe(III)-reducing bacteria, but few are unable to conserve the energy like sulfate-reducing bacteria and clostridium species. The lower standard redox potential of Fe+2 than Pu(V)/Pu(IV), Np(V)/Np(IV), and U(VI)/U(IV) promotes the reduction of aforementioned radionuclide pair by Fe(III)-reducing bacteria either enzymatically or via Fe(II) produced from reduction of Fe(III) oxides by single-electron reduction and formation of cation more amenable to precipitation. Thus reduced natural organic matters NOM, Fe(II), and S(II) have proved to be potential reductants under typical suboxic and anoxic environments (Loughlin et al. 2011).

The most critical issue for all biotransformation is how stable the transformation is. The stability of reduced species for the success of remediation has been studied in terms of the long-term biocycling behavior of radionuclides in natural and engineered environments. For example, complexing agents such as ethylenediaminetetraacetate (EDTA), diethylenetriaminepentaacetic acid (DTPA), oxalate, citrate, humic acid, and fulvic acids promote dissolution of Tc(IV) oxides with the exposure to oxygen or changes in pH. Similarly, reoxidation of U(IV) in sediments with natural microbial communities present in Fe(III) minerals (Sani et al. 2005; Ginder-Vogel et al. 2006; Spycher et al. 2011); manganese oxides (Fredrickson et al. 2002; Wang et al. 2013); organic ligands such as citrate and EDTA, under anaerobic conditions (Luo and Gu 2011); and microbially generated bicarbonate, even under bioreducing conditions (Wan et al. 2005, 2008). Nitrate-reducing bacteria appear to be particularly important in mediating the reoxidation of U(IV) by nitrate (Wilkins et al. 2007).

3 Prospects and Challenges

Remediation of radionuclides through microbial treatment has numerous advantages including being eco-friendly, specificity, adaptability, recyclability of bio-products, etc. Major drawback to this method is slow speed of process and difficulty to control, which accelerates the accumulation of pollutants in the environment with time that can be potentially hazardous. Additionally long-term sustainability of microbial remediation is a question of great importance. With rapid radionuclide accumulation, upgradation of existing bioremediation processes to commercial level by making it faster, recyclable, and controlled is a major challenge.

Polluted environment often contains more than one metal as in the radioactive waste. Therefore, modification in the outer membrane proteins of bacteria with potential bioremediation properties for improving metal binding capacity of microbes may be needed. To large-scale remediation, even complex and combinatorial approaches have been developed to manipulate bacterial genetics in order to design multi-metal resistant bacteria strains. Genetically engineered microorganism is the term used to develop organisms with altered genetic material using recombinant DNA technology. This process is helpful to generate a character-specific efficient strain for biotransformation so that it can be used under various complex environmental conditions. For instance, the gene designated phoK responsible for alkaline phosphatase activity was extracted from Sphingomonas sp. strain BSAR-1. Later it was cloned and overexpressed in Escherichia coli strain BL21(DE3) (Nilgiriwala et al. 2008) and in the radio-resistant bacterium Deinococcus radiodurans (Kulkarni et al. 2013), respectively. The precipitation of >90 % of input uranium has been achieved by recombinant strain in less than 2 h from alkaline waste solutions. Deino-PhoK cell was even not affected by the presence of Cs and Sr. However, for utilizing GEM, sustaining the recombinant bacteria population under various environmental conditions and competition from native bacterial populations are major obstacles reported (Dixit et al. 2015).

Another approach is formation and development of biofilm synthetically. Biofilm is an agglomeration of microbial cells whose thickness is extremely variable ranging from single-cell monolayer to thick mucus microcolonies of microbes held together by extracellular polymeric substances (EPS). The bEPS from Shewanella sp. (Cao et al. 2011) and Desulfovibrio desulfuricans G20 (Beyenal et al. 2004) have been successfully utilized for immobilization of uranium (Fig. 2).

Fig. 2
figure 2

Possible mechanisms of U immobilization and remobilization using SRB biofilms grown on iron-bearing surface. Reprinted (adapted) with permission from Marsili et al. (2007), American Chemical Society

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