**2. Treatment and reclamation of uranium**

#### **2.1. Physical-chemical treatment**

tivity, fission products render the waste matrix in which they are detected highly radiotoxic and therefore hazardous [2]. Some of these elements are easily taken up by plants and other animal life forms upon reaching the environment [3]. As an example, the divalent cation

Calcium is responsible for the bone's structural integrity and strength and is therefore an

Currently, there are about 438 nuclear power plants in operation in 31 countries around the world providing about 14% of the world's primary energy needs [4]. The slow progression towards wider application of nuclear energy technology worldwide is mainly due to concerns over long term radiation contamination, reactor accidents such as those which occurred in Chernobyl and Fukushima, and the possibility of proliferation of atomic bomb making

Nuclear power generation is not the only source of potential radioactive pollution. Other activities such as nuclear weapon testing, radioisotope manufacturing, and biomedical research have also contributed significant amounts of radioactive waste into the environment [5]. However, most radioactive waste originating from medical and radioisotope manufactur‐ ing facilities is predominantly organic and therefore can be easily degraded [6]. Radioactive waste from the power generation industry which is identified as High Level Waste (HLW), on the other hand, consists of a higher proportion of non-degradable metallic elements such as uranium and fission products, and transuranic elements. The waste tends to be "hot" (highly

radiotoxic) to living organisms and therefore requires pre-treatment before disposal.

This chapter presents recent findings from research aimed at developing environmentally friendly treatment processes for radioactive waste. Attempts have recently been made to treat components of high level radioactive waste (HLW) prior disposal in specifically engineered facilities by immobilizing and extracting the radioactive elements in the waste using a combination of biological reduction and biosorption of the toxic metallic elements. In the case of U(VI), the reduced element is easily extracted either by precipitation/deposition on cell surfaces or removed by a biologically mitigated ion exchange process using live cells of

Uranium is found in the environment in many forms including as an oxide, organic or inorganic complex, and rarely as a free metallic ion. Free elemental uranium primarily exists in higher oxidation states typically bound to oxygen. In the aqueous phase, cationic uranium readily combines with oxygen to form oxy-cations of uranium (uranyl ions) which are highly

The toxicity of uranium compounds is closely related to its mobility. That is, the most soluble of the uranium species are associated with acute toxicity in organisms [7]. The highly soluble

2+

) is highly soluble in water.

) is less soluble and therefore

mobile and highly reactive. For example U(VI) in the form of (UO2

Whereas, the reduced form U(IV), existing as uraninite (UO2

) is easily incorporated into bone tissue because its chemical properties

) which is a critical component of the mammalian bone structure.

strontium-90 (90

bacteria.

**1.2. Uranium compound toxicity**

represents a lower risk in the environment.

resemble calcium (Ca2+

Sr2+

180 Applied Bioremediation - Active and Passive Approaches

essential component of the mammalian diet.

materials to renegade regimes and terrorists.

The most common treatment strategy for uranium and radioactive waste involves the extraction of the radioactive component to reduce the volume of radioactive waste followed by treatment of the bulky nonradioactive waste using conventional methods [8]. Various options have been utilised to achieve extraction by employing a combination of physicalchemical and biological methods. For areas that have already been contaminated, further migration of the pollutants is prevented by using *in situ* treatment options [9]. *Ex situ* pumpand-treat processes have also been attempted but these come with very high operational costs and are in most cases unsustainable [10]. *In situ* remediation technologies are generally the most preferred treatment technologies because they are more economical, do not require transportation of toxic materials that may lead to more spillage in transit, and also cause fewer disturbances on site [11, 12].

#### **2.2. Adsorptive/ion exchange processes**

Separation processes have been utilised to selectively remove cationic species from wastewater streams. Materials such as activated carbon, saw dust, and peat can remove pollutants from water. However, these materials are not selective and therefore may not be effective in removing metallic elements from nuclear waste. Specially designed resins can be utilised to target specific species by manipulating the composition of functional groups on surface of the resin. Several examples of uranium binding ion exchange systems are reported by several researchers [13-15]. Although proven successful on pilot scale, full implementation of ion exchange uranium separation is hindered by high cost. Additionally, the ion exchange resin surfaces are not self-regenerating, and therefore have limited capacity [13].

#### **2.3. Membrane processes**

Conventional membrane systems used in treating uranium includes, nano-pore filtration, ultrafiltration, microfiltration and reverse osmosis [16]. Nano-pore membrane filters have the potential to be used in recovery of radioisotopes from water or gas streams. Membrane technology is now regarded as established technology with predictable and reliable processing capability than most current alternatives. Membranes have become relatively cheap such that there use is no longer regarded uneconomical. In spite of being economically viable membrane processes generate large quantities of used membranes which contribute to the problem of radioactive solid waste from the nuclear industry.

the toxic forms of heavy metal and transitional metal elements in transuranic waste (TRU) to less mobile precipitable forms [27]. Other researchers have found microbial cultures with the

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183

Most of the microorganisms discovered thus far, utilise the energy derived from change in the redox potential or oxidation states of the compounds for metabolic purposes. No organism has yet been shown to utilise the energy derived from radioactive decay for metabolism. But recent research show promise that this scenario is soon to change. For example, recent studies have shown that microorganism may not only resist radiation, but may to a certain degree utilize the radiation for metabolic advantage. One example was illustrated in a recent study in which cultures of melanising fungi from the cold regions utilized ionizing radiation to derive

Biosorption is used to describe the metabolism independent sorption (passive process). In this process uranium-bearing water is brought into contact with either living or dead biomass that possesses abandoned functional groups (carboxyl, hydroxyl, amine, and phosphate group) within the surface layer. The charged group within the cell surface layer is able to interact with ions or charged molecules present in the uranium-bearing water. As a result metal cations

Polymers secreted by many metabolizing microbes can also immobilize metals. Different studies on biosorption demonstrated that uranium biosorption is reversible, species-specific, and depends upon the chemistry and pH of the solution, physiological state of cells as well as the presence of the extracellular soluble polymers (EPS) [29, 30]. In this process desorption and recovery of heavy metals and radionuclides for further reuse is easy [31]. Biosorption of radionuclides to the cell surface and polymer substance is a promising technology for reme‐ diating contaminated waters. However, the effectiveness of this process is highly affected by pH of the solution and saturation of the biosorbent when metal interactive sites are occupied. Studies were done [32, 33] to investigate the biosorption of uranium under acid conditions. It was observed in these studies that biosorption under acidic conditions is not favoured in

hydroxyl groups–SOH), thus indicating poor selectivity of the biosorbent surface against

In order to understand the interaction between the cell and metallic species in wastewater, the cell surface can be thoroughly characterised using Fourier Transform Infrared spectroscopy (FTIR) or Energy-Dispersive X-ray spectroscopy (EDX). In our research group at the University of Pretoria, we were able to demonstrate that different functional groups on the cell wall of

Using autotitration date in the software MINTEQ, four equilibrium states were identified

sulphate reducing bacteria (SRB) were associated with the hydroxyl functional groups.

sulphate reducing bacteria acted as ion exchange sites in different pH ranges.

which were associated with carboxylic functional groups (pKa

8-9), and hydroxyl/amines (pKa

) compete with UO2

2+ for sorption sites (surface

4-5), phosphates (pKa 6-7),

10-12). The most abundant reaction sites in

become electrostatically attracted and bound to the cell surface layer.

several species as at low pH the protons (H+

capability to resist high radiation doses.

metabolic energy [28].

**2.6. Biosorption**

competing ions.

phenolics (pKa

#### **2.4. Chemical extraction**

Chemical extraction processes have mostly been used for remediation processes mostly on land. For example, uranium can be extracted from contaminated soil using sodium carbonate/ bicarbonate or citric acid [17-19]. Although this process effectively removes uranium from soil, it requires a careful balance during application since overloading the system with the acid agent may further migrate uranium in the environment [18, 20]. Certain chemical agents may oxidise other potentially toxic metals posing further risk to the environment. Furthermore, long-term stability of reaction products is of concern. Changing chemical conditions in future could remobilise the metal to its toxic form.

#### **2.5. Biological treatment processes**

Biological methods have been proposed to improve or substitute the conventional physicochemical methods for the remediation of contaminated environments. Unlike organic com‐ pounds, toxic metals cannot be degraded or destroyed but can only be transformed from high oxidation state to lower oxidation state. Microbes can potentially affect the physical and the chemical state of the uranium by altering its speciation, solubility, and sorption properties. Strategies suggested for the removal of metals and radionuclides using appropriate microbes include biosorption, bioaccumulation, bioprecipitation, and bioreduction [21-23].

Biological treatment is based on the prospect of utilising processes already devised by nature in dealing with environmental hazards. During three and a half billion years of evolution, microorganisms have evolved mechanisms to survive in hostile environments and to adapt to changes in the environment [24]. One of the most conserved mechanisms in the living cell is the biochemical pathway for electron-transport through the cytoplasmic membrane to conserve energy through the oxidation of an electron donor and reduction of an electron acceptor such as oxygen. This process has been conserved over billions of years, such that, to this day, all life on earth depends on variants of this pathway [24-26]. Most biochemical processes for degradation and/or detoxification of compounds in the living cell are linked to the above process.

Environmental engineers around the world have undertaken to find ways to tap into the mysteries of nature by diligently studying the action of microorganisms as they adapt to extreme conditions. Lately, microorganisms have been isolated that are capable of reducing the toxic forms of heavy metal and transitional metal elements in transuranic waste (TRU) to less mobile precipitable forms [27]. Other researchers have found microbial cultures with the capability to resist high radiation doses.

Most of the microorganisms discovered thus far, utilise the energy derived from change in the redox potential or oxidation states of the compounds for metabolic purposes. No organism has yet been shown to utilise the energy derived from radioactive decay for metabolism. But recent research show promise that this scenario is soon to change. For example, recent studies have shown that microorganism may not only resist radiation, but may to a certain degree utilize the radiation for metabolic advantage. One example was illustrated in a recent study in which cultures of melanising fungi from the cold regions utilized ionizing radiation to derive metabolic energy [28].

#### **2.6. Biosorption**

**2.3. Membrane processes**

182 Applied Bioremediation - Active and Passive Approaches

**2.4. Chemical extraction**

radioactive solid waste from the nuclear industry.

could remobilise the metal to its toxic form.

**2.5. Biological treatment processes**

the above process.

Conventional membrane systems used in treating uranium includes, nano-pore filtration, ultrafiltration, microfiltration and reverse osmosis [16]. Nano-pore membrane filters have the potential to be used in recovery of radioisotopes from water or gas streams. Membrane technology is now regarded as established technology with predictable and reliable processing capability than most current alternatives. Membranes have become relatively cheap such that there use is no longer regarded uneconomical. In spite of being economically viable membrane processes generate large quantities of used membranes which contribute to the problem of

Chemical extraction processes have mostly been used for remediation processes mostly on land. For example, uranium can be extracted from contaminated soil using sodium carbonate/ bicarbonate or citric acid [17-19]. Although this process effectively removes uranium from soil, it requires a careful balance during application since overloading the system with the acid agent may further migrate uranium in the environment [18, 20]. Certain chemical agents may oxidise other potentially toxic metals posing further risk to the environment. Furthermore, long-term stability of reaction products is of concern. Changing chemical conditions in future

Biological methods have been proposed to improve or substitute the conventional physicochemical methods for the remediation of contaminated environments. Unlike organic com‐ pounds, toxic metals cannot be degraded or destroyed but can only be transformed from high oxidation state to lower oxidation state. Microbes can potentially affect the physical and the chemical state of the uranium by altering its speciation, solubility, and sorption properties. Strategies suggested for the removal of metals and radionuclides using appropriate microbes

Biological treatment is based on the prospect of utilising processes already devised by nature in dealing with environmental hazards. During three and a half billion years of evolution, microorganisms have evolved mechanisms to survive in hostile environments and to adapt to changes in the environment [24]. One of the most conserved mechanisms in the living cell is the biochemical pathway for electron-transport through the cytoplasmic membrane to conserve energy through the oxidation of an electron donor and reduction of an electron acceptor such as oxygen. This process has been conserved over billions of years, such that, to this day, all life on earth depends on variants of this pathway [24-26]. Most biochemical processes for degradation and/or detoxification of compounds in the living cell are linked to

Environmental engineers around the world have undertaken to find ways to tap into the mysteries of nature by diligently studying the action of microorganisms as they adapt to extreme conditions. Lately, microorganisms have been isolated that are capable of reducing

include biosorption, bioaccumulation, bioprecipitation, and bioreduction [21-23].

Biosorption is used to describe the metabolism independent sorption (passive process). In this process uranium-bearing water is brought into contact with either living or dead biomass that possesses abandoned functional groups (carboxyl, hydroxyl, amine, and phosphate group) within the surface layer. The charged group within the cell surface layer is able to interact with ions or charged molecules present in the uranium-bearing water. As a result metal cations become electrostatically attracted and bound to the cell surface layer.

Polymers secreted by many metabolizing microbes can also immobilize metals. Different studies on biosorption demonstrated that uranium biosorption is reversible, species-specific, and depends upon the chemistry and pH of the solution, physiological state of cells as well as the presence of the extracellular soluble polymers (EPS) [29, 30]. In this process desorption and recovery of heavy metals and radionuclides for further reuse is easy [31]. Biosorption of radionuclides to the cell surface and polymer substance is a promising technology for reme‐ diating contaminated waters. However, the effectiveness of this process is highly affected by pH of the solution and saturation of the biosorbent when metal interactive sites are occupied. Studies were done [32, 33] to investigate the biosorption of uranium under acid conditions. It was observed in these studies that biosorption under acidic conditions is not favoured in several species as at low pH the protons (H+ ) compete with UO2 2+ for sorption sites (surface hydroxyl groups–SOH), thus indicating poor selectivity of the biosorbent surface against competing ions.

In order to understand the interaction between the cell and metallic species in wastewater, the cell surface can be thoroughly characterised using Fourier Transform Infrared spectroscopy (FTIR) or Energy-Dispersive X-ray spectroscopy (EDX). In our research group at the University of Pretoria, we were able to demonstrate that different functional groups on the cell wall of sulphate reducing bacteria acted as ion exchange sites in different pH ranges.

Using autotitration date in the software MINTEQ, four equilibrium states were identified which were associated with carboxylic functional groups (pKa 4-5), phosphates (pKa 6-7), phenolics (pKa 8-9), and hydroxyl/amines (pKa 10-12). The most abundant reaction sites in sulphate reducing bacteria (SRB) were associated with the hydroxyl functional groups. Adsorption of the divalent fission product (Sr2+) was inhibited at higher pH, supposedly as a result of increased hydroxylation at the high OH- concentration in solution. Additionally, increased pH could increase the formation of SrII-OH precipitates which is counterproductive to the processes of adsorption to the cells.

over time [36]. Additionally, these processes precipitate metals other than uranium and forms

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185

Reduction of highly toxic and mobile U(VI) to sparingly soluble U(IV) using appropriate microbes has been proposed as a mechanism for preventing the migration of U(VI) with groundwater [37, 38]. The strategy is based on injecting physiological electron donors such as acetate, lactate, ethanol, or glucose to stimulate U(VI) reduction by microbial community native to contaminated aquifers [39]. Microorganisms are known to have evolved biochemical pathway for degradation or transformation of toxic compounds from their immediate environment either for survival or to derive energy by using toxic compounds as electron donors or acceptors [40, 41]. The overall transfer of electrons from the carbon sources to active

> *ne -*

An example of a balanced stoichiometric relationship during U(VI) reduction using propa‐

Microbial reduction of U(VI) was first reported in crude extracts from *Micrococcus lactilyticus* by assaying the consumption of hydrogen which was dependent on the presence of U(VI) [42].

2+ - - <sup>+</sup> UO +1/2CH CH COO +1/2H O UO +1/2CH COO +1/2CO +2H 2 32 2 23 ¾¾® <sup>2</sup> (1)

Insoluble reduced form

Soluble U(VI) as electron sink

U(IV)

insoluble uranyl-phosphate complex on the cell surface.

uranium species can be visualised by the figure below (Figure 1).

Electron flow

**Figure 1.** Electron flow during biological reduction of uranium (VI) to U(IV)

noate as an electron donor is represented by Equation 1 (below):

Organic carbon (CHO, acetate, ethanol, etc)

CO2 + H2O

**2.9. Bioreduction**

Biosorption offers a unique advantage in that the biosorbent media (bacteria) is self-regener‐ ating and can be safely disposed after expiry. Apart from the uranium species, the biosorbent can remove a range of other toxic heavy metals from the wastewater without creating haz‐ ardous sludge at costs much lower than conventionally used ion exchange systems. Regener‐ ation of the biosorbent and concentration of the metal solution for eventual recovery further increase the cost effectiveness of the process.

#### **2.7. Bioaccumulation**

Bioaccumulation is an active process wherein metals are taken up into living cells and sequestrated intracellularly by complexing with specific metal-binding components or by precipitation. Intracellular accumulation of metals occurs among all classes of microorganisms by an energy-dependent transport system. Unlike metabolically essential metals such as Fe, Cu, Zn, Co, and Mn, which accumulates intracellularly via energy transport system, intracel‐ lular uranium sequestration is attributed to non-specific transport system mainly due to increased membrane permeability resulting from uranium toxicity in the living cell [34]. Therefore, intracellular accumulation of uranium is considered as metabolism-independent process. The major drawbacks associated with the use of active uptake systems is the require‐ ment of metabolically active cells and also the challenge in metal desorption and recovery [35]. Specifically, the cells will need to be destroyed to release the metal either by lysis or by incineration. Therefore, in this case, the medium for the uptake of metals cannot be reused.

#### **2.8. Bioprecepitation**

Bioprecipitation also known as biocrystallization or biomineralization is the process by which metals and radionuclides can be precipitated with microbially generated ligands such as phosphate, sulphide or oxalate [35]. In these processes bacteria interact strongly with metal ions and concentrate them, eventually generating carbonates and hydroxide minerals at the surface of the cell. Macaskie et al. [35] investigated *Citrobacter sp*. accumulation of heavy deposits of UO2 2+-phosphate, derived from enzymatically liberated phosphate ligand. Cells showed no saturation constrains and it could accumulate several times their own weight of precipitated metal. The above method showed that metal deposition occurs via an initial nucleation pathway involving phosphate groups localized within certain cell-surface lipopo‐ lysaccharides (LPS). Accumulation of metal-phosphate complex within the LPS was suggested to prevent fouling of the cell surface by the accumulated precipitate. The limitations of method during application in an industrial process could be similar to those encountered in biosorp‐ tion. Firstly, the process is hindered by the formation of negatively charged uranyl carbonate complexes, U(VI)–CO3 , arising from microbial metabolism of the carbon source under anae‐ robic conditions and over time the U(VI) carbonate complex formed stimulate U(VI) oxidation over time [36]. Additionally, these processes precipitate metals other than uranium and forms insoluble uranyl-phosphate complex on the cell surface.

#### **2.9. Bioreduction**

Adsorption of the divalent fission product (Sr2+

to the processes of adsorption to the cells.

184 Applied Bioremediation - Active and Passive Approaches

increase the cost effectiveness of the process.

**2.7. Bioaccumulation**

**2.8. Bioprecepitation**

deposits of UO2

complexes, U(VI)–CO3

2+

result of increased hydroxylation at the high OH- concentration in solution. Additionally, increased pH could increase the formation of SrII-OH precipitates which is counterproductive

Biosorption offers a unique advantage in that the biosorbent media (bacteria) is self-regener‐ ating and can be safely disposed after expiry. Apart from the uranium species, the biosorbent can remove a range of other toxic heavy metals from the wastewater without creating haz‐ ardous sludge at costs much lower than conventionally used ion exchange systems. Regener‐ ation of the biosorbent and concentration of the metal solution for eventual recovery further

Bioaccumulation is an active process wherein metals are taken up into living cells and sequestrated intracellularly by complexing with specific metal-binding components or by precipitation. Intracellular accumulation of metals occurs among all classes of microorganisms by an energy-dependent transport system. Unlike metabolically essential metals such as Fe, Cu, Zn, Co, and Mn, which accumulates intracellularly via energy transport system, intracel‐ lular uranium sequestration is attributed to non-specific transport system mainly due to increased membrane permeability resulting from uranium toxicity in the living cell [34]. Therefore, intracellular accumulation of uranium is considered as metabolism-independent process. The major drawbacks associated with the use of active uptake systems is the require‐ ment of metabolically active cells and also the challenge in metal desorption and recovery [35]. Specifically, the cells will need to be destroyed to release the metal either by lysis or by incineration. Therefore, in this case, the medium for the uptake of metals cannot be reused.

Bioprecipitation also known as biocrystallization or biomineralization is the process by which metals and radionuclides can be precipitated with microbially generated ligands such as phosphate, sulphide or oxalate [35]. In these processes bacteria interact strongly with metal ions and concentrate them, eventually generating carbonates and hydroxide minerals at the surface of the cell. Macaskie et al. [35] investigated *Citrobacter sp*. accumulation of heavy

showed no saturation constrains and it could accumulate several times their own weight of precipitated metal. The above method showed that metal deposition occurs via an initial nucleation pathway involving phosphate groups localized within certain cell-surface lipopo‐ lysaccharides (LPS). Accumulation of metal-phosphate complex within the LPS was suggested to prevent fouling of the cell surface by the accumulated precipitate. The limitations of method during application in an industrial process could be similar to those encountered in biosorp‐ tion. Firstly, the process is hindered by the formation of negatively charged uranyl carbonate

robic conditions and over time the U(VI) carbonate complex formed stimulate U(VI) oxidation


, arising from microbial metabolism of the carbon source under anae‐

) was inhibited at higher pH, supposedly as a

Reduction of highly toxic and mobile U(VI) to sparingly soluble U(IV) using appropriate microbes has been proposed as a mechanism for preventing the migration of U(VI) with groundwater [37, 38]. The strategy is based on injecting physiological electron donors such as acetate, lactate, ethanol, or glucose to stimulate U(VI) reduction by microbial community native to contaminated aquifers [39]. Microorganisms are known to have evolved biochemical pathway for degradation or transformation of toxic compounds from their immediate environment either for survival or to derive energy by using toxic compounds as electron donors or acceptors [40, 41]. The overall transfer of electrons from the carbon sources to active uranium species can be visualised by the figure below (Figure 1).

**Figure 1.** Electron flow during biological reduction of uranium (VI) to U(IV)

An example of a balanced stoichiometric relationship during U(VI) reduction using propa‐ noate as an electron donor is represented by Equation 1 (below):

$$\text{UO}\_2^{2+} + 1/2\text{CH}\_3\text{CH}\_2\text{COO}^\* + 1/2\text{H}\_2\text{O} \longrightarrow \text{UO}\_2 + 1/2\text{CH}\_3\text{COO}^\* + 1/2\text{CO}\_2 + 2\text{H}^+ \tag{1}$$

Microbial reduction of U(VI) was first reported in crude extracts from *Micrococcus lactilyticus* by assaying the consumption of hydrogen which was dependent on the presence of U(VI) [42]. To date, U(VI) reduction capability has been identified in more than 25 species of phyloge‐ netically diverse prokaryotes. Examples of these are the mesophilic sulphate-reducing bacteria (*Desulfovibro sp*.) [43], Fe(III)-reducing bacteria (*Geobacter* and *Shawanella sp*.) [44], fermentative bacteria from *Clostridium sp*., [45], *Acidotolerant bacteria* [46], as well as *Myxobacteria sp.* [47]. Some of these organisms have been reported to conserve energy for growth from U(VI) reduction [43, 47], while others reduce U(VI) without energy gain [45, 48, 49]. Researchers such as Lloyd [50], and Wade and Di Christina [51], have demonstrated the importance of dissim‐ ilatory metal-reducing bacteria (DMRB) in reducing toxic form of uranium (U), iron (Fe), manganese (Mn), and other toxic metals. The unique physiological property of DMRD is that they are obligate anaerobes that are capable of utilising (Fe(III) and Mn(IV) oxide) as terminal electron acceptors. In the presence of the radionuclides (U(VI) and Tc(VII)), these could be mostly bioavailable since Fe(III) and Mn(IV) exist as insoluble hydroxides in the natural pH range.

presence of acetate or pyruvate as electron donors [57]. Other researchers observed a partial increase in the U(VI) reduction activity in the presence of either lactate or pyruvate, suggesting

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Genome sequence of *Geobacter sulfurreducens* revealed putative open reading frames (ORFs) for 73 multiheme c-type cytochromes [58]. The genes responsible for U(VI) reduction include, triheme periplasmic cytochrome-*c7*, diheme periplasmic cytochrome, and others [59]. Mutations have been constructed in the number of genes for these proteins and it was observed that each of the mutant constructed from a number of genes negatively affected Fe(III) reduction rates with acetate as electron donor. However, analysis over 15 cyto‐ chromes mutants of *G. sulfurreducens* showed that there was no good correlation between effects on reduction rates of Fe(III) and U(VI) [46]. Interestingly, although this organism is proficient at reducing a broad range of extracellular Fe(III) and Mn(IV) minerals, and UO2

+

exiting in the spent fuel nuclear waste [60]. The authors suggested that the enzyme system responsible for uranium reduction in *G. sulfurreducens* is specific for hexavalent actinides and is capable of transferring one electron to an actinyl ion, and the instability of the

Among the species that conserve metabolic energy from dissimilatory respiration utilising U(VI) an electron sink is *Shewanella oneidensis* MR-1 [61]. *S. putrefaciens* cells cultures under Fe(II) lost their orange colour indicating a major decrease in c-type cytochrome content [61]. The interpretation of these observations was that cytochromes were involved in the transfer of electrons to the terminal electron acceptor or were the terminal reductases. Mutant analysis implicated the nitrate reductase in U(VI) reduction because of the simultaneous

mutant studies have implicated other proteins and cytochromes to be involved in metal reduction and a model for electron transfer was proposed. The function of these electron carriers for U(VI) reduction was only recently evaluated as a part of the analysis of global

By assay of mutants, several proteins including the one involved in menaquinone biosynthesis, decaheme outer membrane cytochrome, a periplasmic decaheme cytochrome, outer mem‐ brane protein, and a tetraheme cytochrome were all shown to be needed for optimal U(VI) reduction [55]. The multiple pathways for electron delivery to U(VI) available in *Shawanella* are associated with the capability of U(VI) reduction with lactate as an electron donor in mutants lacking one or more of the above electron transfer components [55]. Comparison of

to wild-type cells [61]. This result indicate that U(VI) reduction is not eliminated by any of the single mutants analysed and also supports the hypothesis that uranium reductase are likely

showed accumulation predominantly in the periplasm versus the deposition of UO2

) deposition by mutants lacking outer membrane decaheme *c*-type cytochromes

reduction observed in the absence of these reductase [51]. Further

, the reduced species of neptunyl (NpO2

2+

187

2+)

external

the presence of additional proteins capable of metal reduction [54].

it was observed to be inefficient in reducing, NpO2


transcriptional responses to U(VI) [55].

resulting U(V) then generates U(IV) via disproportionation.

*2.10.2. Geobacter reductase*

*2.10.3. Shawanella reductase*

loss of U(VI) and NO3

uraninite (UO2

#### **2.10. Enzymatic U(VI) reduction**

Members of genera *Shewanella* [43], *Desulfovibro* [52], *Clostridium* [45], *Geobacter* [50], *Thermus* [53], *Pyrobaculum* and *Desulfosporosinus* [54], display U(VI) reduction activity. The two approaches (ie. biochemical and genetic) are responsible for identifying U(VI) reductases in DMRB [46, 51, 52, 54, 55]. The mechanism by which Fe(III)-reducing bacteria (FeRB) transfer electrons to insoluble Fe(III) oxides during anaerobic growth have been extensively studied in *Shewanella* and *Geobacter* species [50]. In these organisms, an electron transfer chain containing c-type cytochromes is thought to pass through the periplasmic and terminate at the outer membrane facilitating electron transfer to the extracellular solid phase [50]. U(IV) is reported to precipitate in the periplasm and outside of both gram-negative and gram-positive cells suggesting that U(VI) complexes do not generally have access to intracellular enzyme. Thus, imply that the best candidates for the reductases would be electron carrier proteins or enzymes exposed to the outside of the cytoplasmic membrane, within the periplasm, and/or in the outer membrane. Several purified c-type cytochromes display U(VI) reductase activity in vitro. *Shewanella* and *Geobacter* has been reported to enzymatically reduce U(VI) to U(IV) via a respiratory process that conserve energy to supports their anaerobic growth [52].

#### *2.10.1. Desulfovibrio reductase*

The enzymatic system responsible for U(VI) reduction, tetraheme or periplasmic cytochrome*c3* was characterize in the cell-free extract of the sulphate-reducing bacterium *Desulfovibrio vulgaris* in the presence of hydrogenase using H2 gas as a physiological electron donor [43, 52]. The involvement of tetraheme cytochrome-*c3* was confirmed in the whole-cell and results on whole-cell showed that cytochrome-*c3* was oxidized during U(VI) reduction, but not during sulphate reduction [56]. Further evidence that cytochrome-*c3* was biologically important for *Desulfovibrio* U(VI) reduction came from the impairment of this process when a mutation was constructed in a related strain that eliminated the homologous cytochrome [54]. U(VI) reduction by the mutant *D. desulfuricans* strain was inhibited by at least 90% in the presence of H2 (gas) as an electron donor and partially impaired (inhibition between 50-70%) in the presence of acetate or pyruvate as electron donors [57]. Other researchers observed a partial increase in the U(VI) reduction activity in the presence of either lactate or pyruvate, suggesting the presence of additional proteins capable of metal reduction [54].

#### *2.10.2. Geobacter reductase*

To date, U(VI) reduction capability has been identified in more than 25 species of phyloge‐ netically diverse prokaryotes. Examples of these are the mesophilic sulphate-reducing bacteria (*Desulfovibro sp*.) [43], Fe(III)-reducing bacteria (*Geobacter* and *Shawanella sp*.) [44], fermentative bacteria from *Clostridium sp*., [45], *Acidotolerant bacteria* [46], as well as *Myxobacteria sp.* [47]. Some of these organisms have been reported to conserve energy for growth from U(VI) reduction [43, 47], while others reduce U(VI) without energy gain [45, 48, 49]. Researchers such as Lloyd [50], and Wade and Di Christina [51], have demonstrated the importance of dissim‐ ilatory metal-reducing bacteria (DMRB) in reducing toxic form of uranium (U), iron (Fe), manganese (Mn), and other toxic metals. The unique physiological property of DMRD is that they are obligate anaerobes that are capable of utilising (Fe(III) and Mn(IV) oxide) as terminal electron acceptors. In the presence of the radionuclides (U(VI) and Tc(VII)), these could be mostly bioavailable since Fe(III) and Mn(IV) exist as insoluble hydroxides in the natural pH

Members of genera *Shewanella* [43], *Desulfovibro* [52], *Clostridium* [45], *Geobacter* [50], *Thermus* [53], *Pyrobaculum* and *Desulfosporosinus* [54], display U(VI) reduction activity. The two approaches (ie. biochemical and genetic) are responsible for identifying U(VI) reductases in DMRB [46, 51, 52, 54, 55]. The mechanism by which Fe(III)-reducing bacteria (FeRB) transfer electrons to insoluble Fe(III) oxides during anaerobic growth have been extensively studied in *Shewanella* and *Geobacter* species [50]. In these organisms, an electron transfer chain containing c-type cytochromes is thought to pass through the periplasmic and terminate at the outer membrane facilitating electron transfer to the extracellular solid phase [50]. U(IV) is reported to precipitate in the periplasm and outside of both gram-negative and gram-positive cells suggesting that U(VI) complexes do not generally have access to intracellular enzyme. Thus, imply that the best candidates for the reductases would be electron carrier proteins or enzymes exposed to the outside of the cytoplasmic membrane, within the periplasm, and/or in the outer membrane. Several purified c-type cytochromes display U(VI) reductase activity in vitro. *Shewanella* and *Geobacter* has been reported to enzymatically reduce U(VI) to U(IV) via a

respiratory process that conserve energy to supports their anaerobic growth [52].

The enzymatic system responsible for U(VI) reduction, tetraheme or periplasmic cytochrome*c3* was characterize in the cell-free extract of the sulphate-reducing bacterium *Desulfovibrio*

52]. The involvement of tetraheme cytochrome-*c3* was confirmed in the whole-cell and results on whole-cell showed that cytochrome-*c3* was oxidized during U(VI) reduction, but not during sulphate reduction [56]. Further evidence that cytochrome-*c3* was biologically important for *Desulfovibrio* U(VI) reduction came from the impairment of this process when a mutation was constructed in a related strain that eliminated the homologous cytochrome [54]. U(VI) reduction by the mutant *D. desulfuricans* strain was inhibited by at least 90% in the presence

(gas) as an electron donor and partially impaired (inhibition between 50-70%) in the

gas as a physiological electron donor [43,

range.

**2.10. Enzymatic U(VI) reduction**

186 Applied Bioremediation - Active and Passive Approaches

*2.10.1. Desulfovibrio reductase*

of H2

*vulgaris* in the presence of hydrogenase using H2

Genome sequence of *Geobacter sulfurreducens* revealed putative open reading frames (ORFs) for 73 multiheme c-type cytochromes [58]. The genes responsible for U(VI) reduction include, triheme periplasmic cytochrome-*c7*, diheme periplasmic cytochrome, and others [59]. Mutations have been constructed in the number of genes for these proteins and it was observed that each of the mutant constructed from a number of genes negatively affected Fe(III) reduction rates with acetate as electron donor. However, analysis over 15 cyto‐ chromes mutants of *G. sulfurreducens* showed that there was no good correlation between effects on reduction rates of Fe(III) and U(VI) [46]. Interestingly, although this organism is proficient at reducing a broad range of extracellular Fe(III) and Mn(IV) minerals, and UO2 2+ it was observed to be inefficient in reducing, NpO2 + , the reduced species of neptunyl (NpO2 2+) exiting in the spent fuel nuclear waste [60]. The authors suggested that the enzyme system responsible for uranium reduction in *G. sulfurreducens* is specific for hexavalent actinides and is capable of transferring one electron to an actinyl ion, and the instability of the resulting U(V) then generates U(IV) via disproportionation.

#### *2.10.3. Shawanella reductase*

Among the species that conserve metabolic energy from dissimilatory respiration utilising U(VI) an electron sink is *Shewanella oneidensis* MR-1 [61]. *S. putrefaciens* cells cultures under Fe(II) lost their orange colour indicating a major decrease in c-type cytochrome content [61]. The interpretation of these observations was that cytochromes were involved in the transfer of electrons to the terminal electron acceptor or were the terminal reductases. Mutant analysis implicated the nitrate reductase in U(VI) reduction because of the simultaneous loss of U(VI) and NO3 reduction observed in the absence of these reductase [51]. Further mutant studies have implicated other proteins and cytochromes to be involved in metal reduction and a model for electron transfer was proposed. The function of these electron carriers for U(VI) reduction was only recently evaluated as a part of the analysis of global transcriptional responses to U(VI) [55].

By assay of mutants, several proteins including the one involved in menaquinone biosynthesis, decaheme outer membrane cytochrome, a periplasmic decaheme cytochrome, outer mem‐ brane protein, and a tetraheme cytochrome were all shown to be needed for optimal U(VI) reduction [55]. The multiple pathways for electron delivery to U(VI) available in *Shawanella* are associated with the capability of U(VI) reduction with lactate as an electron donor in mutants lacking one or more of the above electron transfer components [55]. Comparison of uraninite (UO2 ) deposition by mutants lacking outer membrane decaheme *c*-type cytochromes showed accumulation predominantly in the periplasm versus the deposition of UO2 external to wild-type cells [61]. This result indicate that U(VI) reduction is not eliminated by any of the single mutants analysed and also supports the hypothesis that uranium reductase are likely non-specific, and that low potential electron donors are present in both the periplasm and outer membrane. It remains to be determined whether the mutants altered for U(VI) reduction are similarly affected in their ability to use U(VI) as terminal electron acceptor for growth.

#### **2.11. Permeable reactive barriers**

Waste from power generation and fuel process facilities contains high levels of uranium and transuranic elements. This type of waste, classified as high level waste (HLW) or transuranic waste (TU), is usually solidified in a concrete or bitumen before it is stored in specially engineered facilities above the ground. The chances of environmental contamination from such facilities are slight. However, most of the voluminous intermediate level and low level waste (ILW and LLW) can be packed and stored underground. The underground storage facilities pose a high risk of groundwater contamination. Where contamination has actually occurred, pump-and-treat processes are utilised to intercept the polluted groundwater for treatment above ground. The water can be treated using chemicals or using biological reactors and the clean effluent is returned to the aquifer. For toxic metals, chemical agents may be added followed by precipitation to reclaim the metals [62]. The chemical reduction process utilizes toxic reducing agents that produce toxic sludge requiring further treatment before disposal into natural waters. Biological processes have been proposed for the pump-and-treat process, but this does not eliminate the problem of disposal of the product of the precipitation stage. Several techniques for installing a biological barrier have been attempted such as construction of semi-porous walls which require a fair amount of excavation (Figure 2), injection of nutrients to encourage the growth of certain types or native species in the environment (a form of bioaugmentation), and inoculation of a region down gradient of a pollutant with specialized cultures of bacteria. Molecular *in situ* bioremediation, the process of introducing new genetic material in native species, has not been put to practice anywhere apart from small scale experimental projects on petrochemical pollution [63].

*Insitu* bioremediation techniques using permeable reactive barriers (PRBs) have been used to treat organic pollutants that can be completely mineralized to safe products, carbon dioxide (CO2 ) and water (H2 O), by microorganisms in the environment. A wide range of toxic recalci‐ trant organic compounds have been treated this way. For example, case studies have been reported on the treatment of petrochemical pollutants [i.e., benzene, toluene, ethylbenzene, xylene, polychlorinated biphenyls (PCBs), and polycyclic aromatic hydrocarbons (PAHs)], and agricultural pollutants: methyl-tert-butyl-ether (MTBE) and its congeners [7, 63].

The decreasing concentration of U(VI) across the barrier is envisioned if barrier is inoculated with U(VI) reducing bacterial species. In the case of U(VI) reduction across a barrier system, we hope to utilise U(VI) as an electron sink in a dissimilatory respiration process in which the organisms introduced in the barrier (*Xa* ) will require U(VI) to optimise their growth. If the organisms require U(VI) as a growth limiting electron sink, their survival away from the barrier zone will be limited. This will prevent increased microbial counts in the aquifer water if the aquifer downstream of the direction of flow is utilised as a drinking water supply source.

the conditions in the soil change one day in future, the metal may be remobilized to its chemically toxic and mobile state and migrate down gradient to further contaminate ground‐ water and surface water resources. The alternative is to let the reduced form migrate without precipitating in the aquifer medium. This can then be removed by a pump-and-treat method

**Figure 2.** Theoretical representation of the microbial permeable reactive barrier system as an intervention for U(VI) pollution in an unconfined aquifer system. The graph shows the U(VI) concentration and biomass propagation under

hydroxide precipitates of reduction products. The number of complexed hydroxyl ions, *n*, will depend on the charge

hydroxyl ions, *n*, will depend on the charge on the uraninite group UxOyn+.

optimum operation conditions. U = uranium (VI) concentration, Xa

Oy n+ .

removed by a pump-and-treat method as described above.

**3. Determination of U(VI)** 

Figure 2. Theoretical representation of the microbial permeable reactive barrier system as an intervention for U(VI) pollution in an unconfined aquifer system. The graph shows the U(VI) concentration and biomass propagation under optimum operation conditions. U = uranium (VI) concentration, Xa = concentration of active biomass, and U*x*O*y*(OH)n = hydroxide precipitates of reduction products. The number of complexed

Bioremediation of Radiotoxic Elements under Natural Environmental Conditions

http://dx.doi.org/10.5772/56909

189

= concentration of active biomass, and U*<sup>x</sup>*

O*<sup>y</sup>* (OH)n =

*In situ* bioremediation techniques using permeable reactive barriers (PRBs) have been used to treat organic pollutants that can be completely mineralized to safe products, carbon dioxide (CO2) and water (H2O), by microorganisms in the environment. A wide range of toxic recalcitrant organic compounds have been treated this way. For example, case studies have been reported on the treatment of petrochemical pollutants [i.e., benzene, toluene, ethylbenzene, xylene, polychlorinated biphenyls (PCBs), and polycyclic aromatic hydrocarbons (PAHs)], and agricultural pollutants: methyl-tert-butyl-ether (MTBE) and its congeners [7, 63].

The decreasing concentration of U(VI) across the barrier is envisioned if barrier is inoculated with U(VI) reducing bacterial species. In the case of U(VI) reduction across a barrier system, we hope to utilise U(VI) as an electron sink in a dissimilatory respiration process in which the organisms introduced in the barrier (*Xa*) will require U(VI) to optimise their growth. If the organisms require U(VI) as a growth limiting electron sink, their survival away from the barrier zone will be limited. This will prevent increased microbial counts in the aquifer water if the aquifer downstream of the direction of flow is utilised as a drinking water supply

The main limitation of *in situ* bioremediation for treatment of metals such as uranium is that the element is not destroyed but rather trapped in the aquifer matrix in a reduced state. Should the conditions in the soil change one day in future, the metal may be remobilized to its chemically toxic and mobile state and migrate down gradient to further contaminate groundwater and surface water resources. The alternative is to let the reduced form migrate without precipitating in the aquifer medium. This can then be

as described above.

on the uraninite group Ux

source.

The main limitation of *in situ* bioremediation for treatment of metals such as uranium is that the element is not destroyed but rather trapped in the aquifer matrix in a reduced state. Should Bioremediation of Radiotoxic Elements under Natural Environmental Conditions http://dx.doi.org/10.5772/56909 189

non-specific, and that low potential electron donors are present in both the periplasm and outer membrane. It remains to be determined whether the mutants altered for U(VI) reduction are similarly affected in their ability to use U(VI) as terminal electron acceptor for growth.

Waste from power generation and fuel process facilities contains high levels of uranium and transuranic elements. This type of waste, classified as high level waste (HLW) or transuranic waste (TU), is usually solidified in a concrete or bitumen before it is stored in specially engineered facilities above the ground. The chances of environmental contamination from such facilities are slight. However, most of the voluminous intermediate level and low level waste (ILW and LLW) can be packed and stored underground. The underground storage facilities pose a high risk of groundwater contamination. Where contamination has actually occurred, pump-and-treat processes are utilised to intercept the polluted groundwater for treatment above ground. The water can be treated using chemicals or using biological reactors and the clean effluent is returned to the aquifer. For toxic metals, chemical agents may be added followed by precipitation to reclaim the metals [62]. The chemical reduction process utilizes toxic reducing agents that produce toxic sludge requiring further treatment before disposal into natural waters. Biological processes have been proposed for the pump-and-treat process, but this does not eliminate the problem of disposal of the product of the precipitation stage. Several techniques for installing a biological barrier have been attempted such as construction of semi-porous walls which require a fair amount of excavation (Figure 2), injection of nutrients to encourage the growth of certain types or native species in the environment (a form of bioaugmentation), and inoculation of a region down gradient of a pollutant with specialized cultures of bacteria. Molecular *in situ* bioremediation, the process of introducing new genetic material in native species, has not been put to practice anywhere apart from small scale

*Insitu* bioremediation techniques using permeable reactive barriers (PRBs) have been used to treat organic pollutants that can be completely mineralized to safe products, carbon dioxide

trant organic compounds have been treated this way. For example, case studies have been reported on the treatment of petrochemical pollutants [i.e., benzene, toluene, ethylbenzene, xylene, polychlorinated biphenyls (PCBs), and polycyclic aromatic hydrocarbons (PAHs)], and

The decreasing concentration of U(VI) across the barrier is envisioned if barrier is inoculated with U(VI) reducing bacterial species. In the case of U(VI) reduction across a barrier system, we hope to utilise U(VI) as an electron sink in a dissimilatory respiration process in which the

organisms require U(VI) as a growth limiting electron sink, their survival away from the barrier zone will be limited. This will prevent increased microbial counts in the aquifer water if the aquifer downstream of the direction of flow is utilised as a drinking water supply source. The main limitation of *in situ* bioremediation for treatment of metals such as uranium is that the element is not destroyed but rather trapped in the aquifer matrix in a reduced state. Should

agricultural pollutants: methyl-tert-butyl-ether (MTBE) and its congeners [7, 63].

O), by microorganisms in the environment. A wide range of toxic recalci‐

) will require U(VI) to optimise their growth. If the

**2.11. Permeable reactive barriers**

188 Applied Bioremediation - Active and Passive Approaches

experimental projects on petrochemical pollution [63].

(CO2

) and water (H2

organisms introduced in the barrier (*Xa*

Figure 2. Theoretical representation of the microbial permeable reactive barrier system as an intervention for U(VI) pollution in an unconfined aquifer system. The graph shows the U(VI) concentration and biomass propagation under optimum operation conditions. U = uranium (VI) concentration, Xa = concentration of active biomass, and U*x*O*y*(OH)n = hydroxide precipitates of reduction products. The number of complexed hydroxyl ions, *n*, will depend on the charge on the uraninite group UxOyn+. **Figure 2.** Theoretical representation of the microbial permeable reactive barrier system as an intervention for U(VI) pollution in an unconfined aquifer system. The graph shows the U(VI) concentration and biomass propagation under optimum operation conditions. U = uranium (VI) concentration, Xa = concentration of active biomass, and U*<sup>x</sup>* O*<sup>y</sup>* (OH)n = hydroxide precipitates of reduction products. The number of complexed hydroxyl ions, *n*, will depend on the charge on the uraninite group Ux Oy n+ .

*In situ* bioremediation techniques using permeable reactive barriers (PRBs) have been used to treat organic pollutants that can be

process in which the organisms introduced in the barrier (*Xa*) will require U(VI) to optimise their growth. If the organisms require U(VI) as a growth limiting electron sink, their survival away from the barrier zone will be limited. This will prevent increased microbial counts in the aquifer water if the aquifer downstream of the direction of flow is utilised as a drinking water supply

The main limitation of *in situ* bioremediation for treatment of metals such as uranium is that the element is not destroyed but rather trapped in the aquifer matrix in a reduced state. Should the conditions in the soil change one day in future, the metal may be remobilized to its chemically toxic and mobile state and migrate down gradient to further contaminate groundwater and surface water resources. The alternative is to let the reduced form migrate without precipitating in the aquifer medium. This can then be

the conditions in the soil change one day in future, the metal may be remobilized to its chemically toxic and mobile state and migrate down gradient to further contaminate ground‐ water and surface water resources. The alternative is to let the reduced form migrate without precipitating in the aquifer medium. This can then be removed by a pump-and-treat method as described above. completely mineralized to safe products, carbon dioxide (CO2) and water (H2O), by microorganisms in the environment. A wide range of toxic recalcitrant organic compounds have been treated this way. For example, case studies have been reported on the treatment of petrochemical pollutants [i.e., benzene, toluene, ethylbenzene, xylene, polychlorinated biphenyls (PCBs), and polycyclic aromatic hydrocarbons (PAHs)], and agricultural pollutants: methyl-tert-butyl-ether (MTBE) and its congeners [7, 63]. The decreasing concentration of U(VI) across the barrier is envisioned if barrier is inoculated with U(VI) reducing bacterial species. In the case of U(VI) reduction across a barrier system, we hope to utilise U(VI) as an electron sink in a dissimilatory respiration

removed by a pump-and-treat method as described above.

**3. Determination of U(VI)** 

source.
