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Responses exhibited by various microbial groups relevant to uranium exposure
Accepted Manuscript Responses exhibited by various microbial groups relevant to uranium exposure
Nilesh Kolhe, Smita Zinjarde, Celin Acharya PII: DOI: Reference:
S0734-9750(18)30126-5 doi:10.1016/j.biotechadv.2018.07.002 JBA 7277
To appear in:
Biotechnology Advances
Received date: Revised date: Accepted date:
15 February 2018 8 July 2018 9 July 2018
Please cite this article as: Nilesh Kolhe, Smita Zinjarde, Celin Acharya , Responses exhibited by various microbial groups relevant to uranium exposure. Jba (2018), doi:10.1016/j.biotechadv.2018.07.002
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ACCEPTED MANUSCRIPT Responses exhibited by various microbial groups relevant to uranium exposure Nilesh Kolhea,b,Smita Zinjardea*,c,Celin Acharyab*,d a
Institute of Bioinformatics and Biotechnology, Savitribai Phule Pune University, Pune 411007,
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India. b
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Molecular Biology Division, Bhabha Atomic Research Centre, Trombay, Mumbai 400085,
c
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India.
Department of Microbiology, Savitribai Phule Pune University, Pune 411007, India.
d
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Homi Bhabha National Institute, Anushakti Nagar, Trombay, Mumbai 400094, India.
*
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Authors for correspondence
S. Zinjarde address: Institute
of Bioinformatics
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Mailing
and
Biotechnology,
Savitribai Phule
Pune
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University, Pune 411007, India. Phone: + (91) 20 25691333, E-mail: [email protected];
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[email protected], Fax: + (91) 20 25690087 C. Acharya
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Mailing address: Molecular Biology Division, Bhabha Atomic Research Centre, Trombay,
25505326
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Mumbai 400 085, India. Phone: + (91) 22 25592256, E-mail: [email protected], Fax: + (91) 22
ACCEPTED MANUSCRIPT Abstract There is a strong interest in knowing how various microbial systems respond to the presence of uranium (U), largely in the context of bioremediation. There is no known biological role for uranium so far. Uranium is naturally present in rocks and minerals. The insoluble nature of the
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U(IV) minerals keeps uranium firmly bound in the earth’s crust minimizing its bioavailability.
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However, anthropogenic nuclear reaction processes over the last few decades have resulted in
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introduction of uranium into the environment in soluble and toxic forms. Microbes adsorb,
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accumulate, reduce, oxidize, possibly respire, mineralize and precipitate uranium. This review focuses on the microbial responses to uranium exposure which allows the alteration of the forms
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and concentrations of uranium within the cell and in the local environment. Detailed information on the three major bioprocesses namely, biosorption, bioprecipitation and bioreduction exhibited
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by the microbes belonging to various groups and subgroups of bacteria, fungi and algae is
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provided in this review elucidating their intrinsic and engineered abilities for uranium removal.
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The survey also highlights the instances of the field trials undertaken for in situ uranium bioremediation. Advances in genomics and proteomics approaches providing the information on
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the regulatory and physiologically important determinants in the microbes in response to
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uranium challenge have been catalogued here. Recent developments in metagenomics and metaproteomics indicating the ecologically relevant traits required for the adaptation and survival of environmental microbes residing in uranium contaminated sites are also included. A comprehensive understanding of the microbial responses to
uranium can facilitate the
development of in situ U bioremediation strategies. Keywords: Uranium, Microorganisms, Biosorption, Bioprecipitation, Bioreduction, Proteomics, Metaproteomics, Genomics, Metagenomics
ACCEPTED MANUSCRIPT 1. Introduction Anthropogenic activities related to nuclear processes such as mining, fuel processing, weapon production or nuclear accidents have resulted in contamination of the environment with uranium (U) (Garbisu and Alkorta, 2003; Markich, 2002). Uranium, with a density of 19 g cm-3 ,
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is a naturally occurring long-lived radionuclide that prevails in soils, rocks, seas and oceans
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(Gavrilescu et al., 2009). The most common isotopes of U in natural deposits are U-238, U-235
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and U-234 and they occur in the proportions of 99.27%, 0.72% and 0.0056%, respectively. Out
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of the three isotopes, U-238 is considered to be stable and has a half-life of 4.5 billion years (Krane, 1987). The average concentration of uranium in earth’s crust (naturally occurring) is
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reported to be 3 mg kg-1 (Kalin et al., 2005). In river water, uranium concentration ranges from 0.01-6.6 µg l-1 (depending on the strata) while 30 and 3.32 µg l-1 U have been reported in
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groundwater and seawater respectively under natural conditions (Markich, 2002). Very high
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concentrations of uranium have also been reported in some parts of the world as a result of
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various nuclear activities. For example, concentrations as high as 11.7 g l-1 in groundwater and 16 mg g-1 in soil and sediments have been reported at various United States Department of
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Energy (DOE) sites (Riley and Zachara, 1992). Nuclear waste storage sites at Oak Ridge
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National Laboratory, TN, U.S. are also reported to have high (0.8 mg g-1 ) contents of U (Kelly et al., 2007). The chemical guideline value for uranium in drinking water has been increased from 2 μg l-1 in 1998 to 15 μg l-1 in 2004 and subsequently to 30 μg l-1 in 2011 based on human studies from Sweden, Finland and USA (Ansoborlo et al., 2015). Uranium has two major oxidation states, +4 U(IV) and +6 U(VI) (Markich, 2002). The speciation of U depends on factors such as pH, redox potential, temperature, ionic strength, presence of other metals, suspended particles and dissolved CO 2 (Bernhard et al., 1998; Markich,
ACCEPTED MANUSCRIPT 2002; Lu et al., 2013). U(IV) is prevalent in ores in the form of uraninite [UO 2 (s)], that is insoluble and stable under anaerobic conditions. U(VI) on the other hand, prevails in aqueous systems under oxidizing conditions and speciation of this form is dependent on pH. Under acidic conditions (pH < 5), U(VI) exists in the soluble form as UO 2 2+ and is considered to be more toxic
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than U(IV) ions (Suzuki and Banfield, 2004). Hydroxides of uranium that predominate between
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pH 5.0 to 7.0 are less toxic towards aquatic flora and fauna (Markich, 2002; Choppin, 2007).
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Uranyl carbonate species such as UO 2 (CO 3 )2 2- and UO 2 (CO 3 )3 4- are predominant in rivers, ponds, lakes and sea water above circumneutral pH (>7.5) (Konstantinou and Pashalidis, 2004). 235
U) and chemical toxicity. Uranium
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Uranium, exhibits radiological (particularly from
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containing waste released into the environment can accumulate in the soil or move into aquatic systems like surface and groundwater (Gavrilescu et al., 2009). Consequently, uranium is likely
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to be introduced into terrestrial and aquatic food chains (Priest, 2001; Anke et al., 2009). Unlike
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some other metals, uranium does not have any biological role in living organisms. This metal has increased affinity for phosphate groups and binds strongly to phosphorylated peptides (Pardoux
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et al., 2012). In vivo studies have demonstrated that U treatment affects radical scavenging
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activity negatively in lung epithelial cells due to loss of glutathione (an effective antioxidant) and superoxide dismutase (SOD) enzyme (Periyakaruppan et al., 2007). U is known to cause hepatic,
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lung and renal damage in humans (Priest, 2001; Anke et al., 2009). Although many microbes have been suggested to respire uranium (Wade and Christina, 2000; Wagner et al., 2012), the toxicity of uranium in microbes has been established in terms of inhibition of microbial activity (Carvajal et al., 2012), distortion of cell surfaces and loss of cell viability (Sepulveda-Median et al. 2015), abortion of transcriptional and translational processes (Mukherjee et al., 2012), growth arrest and suspension of DNA replication (Park and Jiao, 2014) and oxidative damage (Volesky
ACCEPTED MANUSCRIPT and Holan, 1995). In order to resist uranium toxicity, the microbes employ various mechanisms which have shown potential for uranium immobilization and bioremediation. Uranium contamination poses substantial threat to human, environmental health and safety, therefore demanding a need to address the problem effectively. Physiochemical process
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used for treatment of metal contamination such as membrane filtration, chemical precipitation,
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electrochemical treatment, use of ion-exchange resins, incineration, excavation and extraction
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with organic solvents are often cost-intensive. Moreover, they are ineffective in removing toxic metals from large volumes leading to generation of hazardous by-products. The problems
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associated with physiochemical processes can be overcome by using microbial systems (Wang
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and Chen, 2006). Microorganisms interact with uranium and bring about alterations in the concentrations and/or oxidation states within the cell and their immediate environment thus
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regulating the mobility of uranium in the environment (Macaskie et al., 1992; Merroun and
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Seleneska-Pobell, 2008; Acharya, 2015; Majumder and Wall, 2017). Despite of the toxicity imposed by uranium, native microorganisms have shown tolerance and survival in uranium
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contaminated sites. Such isolates obtained from different parts of the world demonstrating
microorganisms
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significant uranium tolerance are summarized in Table 1. Various strategies adopted by for
persistent
existence
in
uranium
contaminated
environments
include
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mechanisms such as biosorption, biomineralization and biotransformation which can be used for in situ or ex situ uranium bioremediation (Fig. 1). Previously, reviews on uranium have discussed about speciation and bioavailability (Markich, 2002), bioreduction (Wall and Krumholz, 2006; Majumder and Wall, 2017), bacterial and cyanobacterial interactions (Pollmann et al., 2006; Merroun and Seleneska-Pobell, 2008; Acharya and Apte, 2013b) and biogeochemistry and bioremediation of soil ( Gavrilescu et al.,
ACCEPTED MANUSCRIPT 2009; Newsome et al., 2014; Selvakumar et al., 2018). However, to the best of our knowledge, there is no comprehensive recent review detailing uranium interactions with different microbial groups and subgroups of prokaryotes (archaea and bacteria-Gram positive/Gram negative) and
Basidiomycetous Rhodophyta).
and
We
and
filamentous
Zygomycetous
discuss
in
groups
detail three
fungi and
sub-categorized
into
algae-Chlorophyta,
microbial processes
Ascomycetous, Phaeophyta
and
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(fungi-yeasts
important for uranium
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eukaryotes
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bioremediation namely, biosorption, bioprecipitation and bioreduction with respect to various microbial groups (Fig. 1). Subsequently, the role of natural and genetically engineered strains in
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mediating such interactions is discussed. We conclude by detailing the use of ‘omics’ based
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approaches for understanding uranyl interactions at the molecular level. 2. Biosorption of uranium
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Biosorption is a property of live or dead biomass to adsorb and concentrate toxic metals
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from very dilute aqueous solutions including uranium (White and Gadd, 1990; Volesky and Holan, 1995). In the recent years, biosorption has emerged as an important alternative for
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removing radionuclides from aqueous solutions (Lloyd and Macaskie, 2002). This phenomenon is
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often rapid and growth-independent since metabolic processes are not involved. Biomass generated from different biotechnological and agricultural processes can be used as biosorbents
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presenting the economic viability of the biosorption process (Ahluwalia and Goyal, 2007). This method often displays high efficiency and selectivity for metals over a broad range of pH and temperature. A summary of important parameters studied during the process of microbial biosorption of uranium are included in Table 2. Different microbial groups that have been used for biosorption of uranium are detailed in the following sections. Few examples of bioaccumulation have also been included in this section. Bioaccumulation is sometimes referred interchangeably
ACCEPTED MANUSCRIPT with biosorption wherein the accumulated metal is located intracellularly in the metabolically active cells. 2.1. Biosorption of uranium by archaea and bacteria Archaea and bacteria are ubiquitous and often prevail in extreme environments (Battista,
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1997; Selenska-Pobell et al., 2001; Jroundi et al., 2007; Acharya et al., 2012). They have high
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surface to volume ratios and different functional groups (phosphate, carboxyl, amide and
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hydroxyl) present on their surfaces which complex with metals and radionuclides (Vieira and Volesky, 2000; Bader et al. 2018). Archaea and bacteria differ in cell wall characteristics. Unlike
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bacteria, archaea lack peptidoglycan in their cell walls.
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2.1.1. Biosorption of uranium by archaea
The role of archaea belonging to genus Halobacterium in bringing about uranium
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biosorption has been recently demonstrated. Two strains of Halobacterium noricense isolated
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from Waste Isolation Pilot Plant repository were able to associate with uranium in a multistage process. An initial phase of rapid sorption was followed by uranium release and subsequently, a
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slow re-association of uranium with the cells was observed. Carboxylic and phosphoryl groups
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were important in uranyl binding during the first phase of biosorption (Bader et al., 2017; 2018). 2.1.2. Biosorption of uranium by Gram positive bacteria
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Members of the Bacillus, Geobacillus, Clostridium, Streptomyces and Arthrobacter have been shown to function as efficient uranium biosorbents. Bacillus species have been used for biosorption of uranium on several occasions. The site specific complex model (SCM) initially described by Fein et al. (1997) has been used for quantitative prediction of uranyl adsorption onto the cells of Bacillus subtilis under different conditions (Fowle et al., 2000; Gorman-Lewis et al., 2005). Recently, the efficacy of alginate-chitosan microcapsules with immobilized B.
ACCEPTED MANUSCRIPT subtilis in uranium sorption has been demonstrated (Tong, 2017). Proteinaceous surface layer (Slayer) representing the outermost cell envelope component of bacteria has been implicated in uranium biosorption. In Bacillus sphaericus JG-A12 (isolated from uranium mining waste pit), the S-layer was shown to bind uranium via phosphate and carboxyl groups (Raff et al., 2002;
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Merroun et al., 2005). Cells, spores and S-layer of this strain embedded in silica gel displayed
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promising potential for uranium adsorption from wastewater samples (Raff et al., 2002, 2003).
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Extended X-ray absorption fine structure (EXAFS) analysis suggested the coordination of U with phosphate groups in meta-autunite mineral phase (Merroun et al., 2005). Pelosinus sp. strain
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UFO1 [native to Oak Ridge Field Research Centre (FRC), TN, U.S.] showed potential for
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sequestering uranium in two different oxidation states, U(IV) and U(VI) (Ray et al., 2011). Furthermore, the uranium-binding complex (UBC) of strain UFO1 comprising of two S-layer
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domain proteins was shown to bind U(IV) (Thorgersen et al., 2017) in contrast to S-layer of
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Bacillus sphaericus JG-A12 which bound U(VI) (Merroun et al. 2005). Detailed analysis of uranium biosorption in Bacillus sp. (dwc-2) obtained from a radioactive waste disposal site in
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Southwest China revealed the involvement of ion-exchange, complexation and bioaccumulation
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process (Li et al., 2014). Biomass based solid-phase extraction of metals from environmental samples has emerged as a promising technique owing to its simplicity and low application cost
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(Özdemir et al., 2013). Cells of Bacillus mojavensis and Bacillus vallismortis loaded onto multiwalled carbon nanotubes were employed as solid-phase extractors to pre-concentrate uranium from aqueous solutions and lake water samples (Özdemir et al., 2017a, 2017b). In batch process at pH 5.0, multiwalled carbon nanotubes with immobilized B. vallismortis could load up to 25.8 mg U g-1 and could be used for 30 cycles of adsorption and desorption without significant decrease in biosorption efficiency. In yet another study, a thermophilic bacterium, Geobacillus
ACCEPTED MANUSCRIPT thermoleovorans subsp. stromboliensis immobilized on Amberlite XAD-4 ion-exchange resin showed U sorption of 11 mg U g-1 at pH 5.0 (Table 2) (Özdemir and Kilinc, 2012). Ionic Liquids (ILs) are commonly used to concentrate and extract radionuclides from waste streams (Binnemans, 2007; Rout et al., 2012). Certain ILs (1-butyl-3-methylimidazolium
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hexafluorophosphate [BMIM] [PF6 ], N-ethylpyridinium trifluoroacetate [EtPy] [CF 3 COO] and
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N-ethylpyridinium tetrafluoroborate [EtPy] [BF 4 ]) were shown to decrease U biosorption by
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Clostridium sp. due to membrane damage of the cells stressing the need to formulate biocompatible ionic liquids for efficient microbial U recovery (Zhang et al., 2014).
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Several actinomycetes have shown the promising potential for uranium adsorption (Friis
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and Myers-Keith, 1986; Tsuruta, 2004). For example, Streptomyces levoris demonstrated essential properties for biosorption of uranium from acidic solutions which was mediated via
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phosphate groups of teichoic acids harbored in the cell walls. These groups conferred a net
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negative charge on bacterial cells and allowed strong binding of positively charged uranyl ions in pH range of 3.5-6.0 (Tsuruta, 2004). Similarly, the phosphodiester residues of cell wall and
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cytoplasmic fraction of Streptomyces longwoodensis bound uranium with appreciable efficiency
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(Friis and Myers-Keith, 1986). Recent studies on Streptomyces sporoverrucosus dwc-3 isolated from a radioactive waste disposal site in China have revealed the binding of U to amino,
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phosphate and carboxyl groups of the cell walls at pH 3.0 (Li et al., 2016). Various Arthrobacter species included within actinomycetes have been reported for uranium biosorption. Arthrobacter strains are abundantly found in uranium rich environments (Updegraff and Douros, 1972; Miller et al., 1987; Suzuki and Banfield, 2004; Chapon et al., 2012; Kumar et al., 2013a). Arthrobacter ilicis accumulated uranium intracellularly as precipitates within polyphosphate granules (Suzuki and Banfield, 2004). Reduced uranium adsorption was observed in Arthrobacter species G975 in
ACCEPTED MANUSCRIPT the presence of aqueous bicarbonate, a competing ligand for U(VI) (Carvajal et al., 2012). Uranium mobility is governed in aquatic systems by the formation of highly soluble and stable uranyl carbonate complexes namely, UO 2 CO 3 0 , UO 2 (CO 3 )2 2- and UO 2 (CO 3 )3 4- above circumneutral conditions (Langmuir, 1978; Guillaumont et al., 2003). These uranyl species
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interfere with uranium complexation on microbial surfaces. Two more actinomycetes namely
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Amycolatopsis sp. K47 and Brachybacterium sp. G1 have recently been shown to remove
related to these groups of bacteria are detailed in Table 2.
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2.1.3. Biosorption of uranium by Gram negative bacteria
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uranium via biosorption (Celik et al., 2018; Bader et al., 2018). Important biosorption parameters
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Lipopolysaccharide (LPS) and cell wall components play an important role in metal binding in Gram negative bacteria. Phospholipids and phosphate groups of LPS are particularly
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significant for binding of metal ions (Beveridge and Fype, 1985). Different types of Gram
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negative bacteria exhibiting U adsorption are catalogued here. Two members of the family Enterobacteriaceae namely, Citrobacter and Serratia have
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shown prominence for their uranyl sorption abilities (Xie et al., 2008; Kumar et al., 2011).
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Adsorption of U(VI) by Citrobacter freudii was found to be rapid with carboxyl groups playing an important role in uranium binding (Xie et al., 2008). Dead biomass of this bacterium
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displayed better efficiency than live cells indicating a metabolism-independent process of uranium sorption. A strain of Serratia marcescens isolated from sub-surface soil of a uranium ore deposit exhibited high tolerance towards uranium (4 mM) under acidic conditions (pH 3.5). This bacterium removed 92% (21-22 mg l-1 ) and 60-70% (285-335 mg l-1 ) of uranium from 100 µM and 2 mM uranyl solutions, respectively (Kumar et al., 2011). Similarly, an extremely uranium tolerant strain, Cupriavidus metallidurans CH34, resisting up to 30 mM uranium,
ACCEPTED MANUSCRIPT immobilized U(VI) by complexing it with phosphoryl or carboxyl containing biomolecules of LPS layer (Llorens et al., 2012). Uranium adsorption onto bacterial cell surface is largely dependent on pH. Uranium speciation and characteristics of bacterial cell wall change as the function of pH. The uranyl
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content adsorbed onto Shewanella oneidensis MR-1 below pH 5.0 was unaffected by dissolved
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inorganic carbon (DIC) concentrations due to the absence of uranyl carbonate complexes.
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However, at higher pH, the extent of uranium adsorption decreased with increasing concentrations of DIC because of the formation of anionic uranyl-hydroxide-carbonates on bacterial surfaces
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(Sheng and Fein, 2013).
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Members of the genus Pseudomonas (belonging to the Gammaproteobacteria group) have also been studied with respect to uranium biosorption (Hu et al., 1996; Choi and Park, 2005; Choi
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et al., 2009). The live or dead cells of Pseudomonas putida exhibited higher specificity for uranyl
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adsorption when exposed to mixed metal waste solutions comprising of uranium (U), lead (Pb), cadmium (Cd), zinc (Zn) and nickel (Ni) at pH 6.0 (Choi et al., 2009). Compared to soil and clay
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constituents, dead cells of P. putida were more effective in uranium adsorption (Choi and Park,
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2005). The cells of Pseudomonas aeruginosa CSU showed rapid intracellular accumulation of uranium within < 10 s of exposure (Strandberg et al., 1981). The same organism demonstrated
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uranyl adsorption in the presence of transition metals under acidic and neutral conditions equivalent to commercial cation-exchange resins. However, certain cations [ferric iron (Fe 3+) acting as uranium analogues] inhibited binding of U to P. aeruginosa biomass in a significant manner (Hu et al., 1996). Removal of iron from wastewater was therefore proposed as a prerequisite step for efficient utilization of P. aeruginosa CSU biomass for uranium adsorption.
ACCEPTED MANUSCRIPT Uranyl adsorption onto Myxococcus xanthus (belonging to Deltaproteobacteria) was demonstrated to be dependent on culture age, concentration of uranyl ion and pH. Older cultures (72 h) when exposed to 1 mM uranium showed maximal adsorption (2.4 mM of U g-1 of dry biomass) at pH 4.5 (Table 2). Adsorbed uranium was found to be distributed on the cell wall and
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extracellular mucopolysaccharide (Gonzalez-Munoz et al., 1997). Another strain of M. Xanthus
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demonstrated uranium coordination with cell surface-associated organic phosphate groups at very
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low pH (2.0) (Jroundi et al., 2007).
Cyanobacteria are morphologically diverse, oxygenic, photosynthetic, Gram negative
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bacteria commonly present in freshwater, marine and terrestrial habitats (Acharya and Apte
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2013b). They are capable of tolerating, accumulating and detoxifying metals in water bodies thereby altering their mobility and bioavailability (Parker et al., 2000; Pradhan and Rai, 2000;
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Zhou et al., 2004; Acharya et al., 2013). Majority of the uranyl binding studies with
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microorganisms have been performed in the pH range of 1.5 to 6.0 wherein, the uranyl ion (UO 2 2+) and hydroxo complexes like UO 2 (OH)+ and (UO 2 )2 (OH)2 2+ predominate (Markich,
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2002; Choppin, 2007). The inability of several microorganisms to sequester U above pH 7.0 is
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attributed to the repulsion between anionic carbonate complexes of U such as [UO 2 (CO 3 )2 2-] and [UO 2 (CO 3 )3 4-] and negatively charged cell surfaces (Markich, 2002; Konstantinous and
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Pashalidis, 2004; Acharya and Apte, 2013b). However, two marine cyanobacteria have demonstrated the ability to remove significant quantities of uranium from aqueous solutions above pH 7.0 (Acharya et al., 2009, 2012, 2013; Acharya and Apte 2013a). A unicellular marine cyanobacterium, Synechococcus elongatus showed rapid binding (53.5 mg g-1 dry weight of U) within 5-10 min when exposed to 100 µM U supplemented as uranyl carbonate at pH 7.8 (Acharya et al., 2009). Most of the bound uranium complexed with amide and deprotonated
ACCEPTED MANUSCRIPT carboxyl groups of extracellular polysaccharides (EPS) could be released by HCl or EDTA treatment. The organism tolerated 0.5 M NaCl and demonstrated high uranyl loading (2960 µg g1
) from simulated sea water in 4 weeks under continuous replenishment conditions (Acharya et
al., 2013). Immobilized cells of S. elongatus were also effective for U binding under continuous conditions
(Acharya
and
Apte,
2013b).
Regeneration
of
immobilized
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flow-through
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Synechococcus biomass for multiple adsorption-desorption cycles (up to 3 times) was observed
uranium by S.
elongatus,
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above pH 7.0 without any significant loss of U binding (90-92%). In contrast to rapid binding of another marine cyanobacterium, namely Anabaena torulosa
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demonstrated rather slow accumulation of uranium in polyphosphate bodies when the cells were
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exposed to 100 µM uranyl carbonate at pH 7.8 for 24 h (Acharya et al., 2012, 2013a). A naturally abundant bloom-causing cyanobacterium, Microcystis aeruginosa, exhibited potential
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for uranyl adsorption between pH 4.0 to 8.0. Interestingly, uranyl adsorption by this organism
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increased when pH was shifted from 10.0 to 11.0 due to a decrease in intrinsic viscosity of the slime suspension at higher pH (Li et al., 2004).
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2.2. Biosorption of uranium by yeasts and filamentous fungi
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Yeasts and filamentous fungi are lower eukaryotes that can be easily cultivated to produce large quantities of biomass. They are amenable to genetic manipulation and have several
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biotechnological and industrial applications (Volesky and Holan, 1995; Bai and Abraham, 2003; Zinjarde, 2014; Zinjarde et al., 2014). Fermentation processes involving yeasts and fungi result in the generation of large quantities of relatively inexpensive biomass. This biomass has been used for effective biosorption of toxic metals and radionuclides (Kapoor and Viraraghavan, 1995; Wang and Chen, 2006). The role of yeasts and filamentous fungi in biosorption of uranium from aqueous solutions has been discussed in the following sections.
ACCEPTED MANUSCRIPT 2.2.1. Biosorption of uranium by yeasts Easy cultivation on a large scale with high yields and generation of waste biomass from food and beverage industries are some of the advantages of yeasts which present them as promising biosorbents for heavy metal decontamination (Volesky and Holan, 1995). Uranyl
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adsorption by the ascomycetous yeast Saccharomyces cerevisiae has been reported on several
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occasions (Strandberg et al., 1981; Liu et al., 2010; Faghihian and Peyvandi, 2012; Lu et al.,
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2013). Dead cells of S. cerevisiae demonstrated higher metal biosorption capacity compared to the live cells (Soares et al., 2002; Xia et al., 2013). The heat killed cells of S. cerevisiae
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apparently released phosphate due to membrane damage which complexed with uranyl residues
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and formed uranyl phosphate nanoparticles on the cell surface (Wang et al., 2017). Amidoxime groups, known for their selectivity for uranium were successfully grafted on S. cerevisiae
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biomass and used for uranyl adsorption from salt lake brine solutions with good efficiency (Bai
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et al., 2016). Dead biomass of S. cerevisiae with rough surfaces and ‘nano-holes’ provided larger contact area for uranyl sorption as compared to the live cells with smooth surfaces (Wang et al.,
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2017). A marine strain of the ascomycetous yeast, Yarrowia lipolytica isolated from oil-polluted
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seawater (Zinjarde and Pant, 2002) is currently being explored in the context of uranyl removal (Kolhe et al., unpublished data). Preliminary results have demonstrated the ability of this strain
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to sequester 50% of the input uranium (50 µM U) loading up to 37.51 mg U g-1 dry weight of cells at pH 7.5. The cellular localization of the uranyl deposits in the yeast cells was examined by electron microscopy. Scanning electron microscopy (SEM) analysis of the cells exposed to 50 µM uranyl carbonate for 24 h revealed irregular, crustaceous surface in contrast to control, U untreated cells which showed smooth cell surfaces (Figs. 2A and 2B). Transmission electron microscopy (TEM) of U exposed yeast cells revealed electron dense, crystalline needle/spicule
ACCEPTED MANUSCRIPT like deposits on cell surfaces and at intracellular locations (Fig. 2D) as compared to control, U untreated cells which did not show any such crystalline precipitates (Fig. 2C) establishing uranyl association with the yeast cells. Metals interact with different functional groups of proteins, lipids or carbohydrates
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present on cell walls of the biomass (Chen et al., 2007). Chemical modification of these groups
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can modulate biosorption processes. For example, esterification of carboxyl groups and
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methylation of the amino groups in the cell wall of the basidiomycetous yeast, Rhodotorula glutinis by methanol and formaldehyde treatment respectively, enhanced uranyl adsorption as
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compared to the chemically unmodified cells (Bai et al., 2009; 2010). Biomass is usually
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recovered by centrifugation or filtration after adsorption which adds to the treatment costs. ‘Smart biosorbents’ resulting from introduction of magnetic phase into the microbes allowed
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metal removal and biomass recovery by applying external magnetic fields (Wong and Fung,
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1997; Wang et al., 2000; Yavuz et al., 2006; Rao et al., 2013). Magnetically modified biomass of R. glutinis has been successfully utilized for uranium biosorption (Bai et al., 2012).
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Immobilization of microbial biomass imparts mechanical strength, confers chemical stability and
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allows regeneration of the biomass. Calcium-alginate immobilized biomass of R. glutinis showed improved U removal from 0.74 to 17.3 mg g-1 U when the input uranyl concentration was
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increased from 2.5 to 40 mg l-1 , respectively (Bai et al., 2014). Various parameters related to biosorption of uranium by raw and modified biomass of R. glutinis are detailed in Table 2. Recently, another species of Rhodotorula namely, R. mucilaginosa BII-R8 is reported in context of uranium biosorption (U-hydroxides and U-hydroxo-carbonates) by means of carboxyl and phosphate groups of the biomass (Lopez-Fernandez et al., 2018). 2.2.2. Biosorption of uranium by filamentous fungi
ACCEPTED MANUSCRIPT Filamentous fungi have high area to mass ratios and their biomass has been used for the removal of metal ions from the aqueous solutions (Kapoor and Viraraghavan, 1995). Filamentous fungi belonging to the ascomycetous group have been studied for biosorption of uranium as discussed here. Viable and non-viable biomass of Trichoderma harzianum has
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demonstrated better uranyl adsorption capacities than commercial resins such as Dowex-SBR-P
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and IRA-400 (Akhtar et al., 2007). This biomass could be used over five cycles of adsorption-
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desorption thereby presenting its cost-effectiveness for uranyl removal. Calcium alginate immobilized biomass of T. harzianum showed better biosorption efficiency than free cells
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(Akhtar et al., 2009). Higher values of sorption (87-97%) were observed for immobilized cells as
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compared to free cells which demonstrated 40-75% when exposed to 100-400 mg l-1 of input U within 8 h (Akhtar et al., 2009). Similarly, another ascomycetous fungus, Aspergillus fumigatus
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immobilized in calcium alginate, was used for biosorption of uranium with appreciable
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efficiency (Wang et al., 2010). Chemical modification by formaldehyde, methanol and acetic acid enhanced affinity of a mangrove endophytic fungus, Fusarium sp. #ZZF51 towards uranium
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(Chen et al., 2014). Non-viable cells of two air-borne, uranium and acid tolerant fungi,
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Aphanocladium spectabilis and Acremonium minutisporum, showed maximum uranium sorption capacities of 162.1 and 161.5 mg U g-1 of dried biomass, respectively (Gargarello et al., 2008).
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Basidiomycetous fungi have also been implicated for removal of uranium from aqueous solutions. Lentinus sajor-caju is a saprophytic basidiomycetous white-rot fungus that grows easily on carbon sources such as cellulose and produces several extracellular enzymes for bioremediation of xenobiotics compounds (Rodrıguez et al., 2004). The alkali treated biomass of this fungus was found to be more effective in removing uranium than its untreated form (Bayramoğlu et al., 2006). Improved sorption capacity of alkali-treated biomass was apparently
ACCEPTED MANUSCRIPT due to increased availability of binding sites resulting from deacetylation of chitin to chitosan. Another basidiomycetous fungus
Schizophyllum
commune
showed
uranium accumulation
intracellularly within the vacuoles and on the cell walls (Gunther et al., 2014). Unlike other uranyl-microbial interactive studies which demonstrated a predominant role of carboxyl groups
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in uranium complexation (Merroun et al., 2005; Acharya et al., 2009), the phosphate groups
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rather than the carboxyl groups facilitated the uranyl adsorption in S. commune.
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Rhizopus arrhizus, a zygomycetous fungus has also been explored for biosorption of uranium from aqueous solutions (Tsezos and Volesky, 1982; Tsezos, 1983). In an attempt to
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evaluate the competing ion effect, it was observed that aluminiun did not interfere in the
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kinetics of uranyl adsorption in R. arrhizus (Tsezos et al., 1996). Certain products derived from fungi like chitin and chitosan have also shown the ability for biosorption of uranium
Algae
are
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2.3. Biosorption of uranium by algae
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(Sakaguchi et al., 1981; Wang et al., 2009).
photosynthetic,
autotrophic
organisms
abundantly
found
in
aquatic
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environments that can serve as inexpensive biosorbent material. Biosorption of uranium by
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different groups of algae belonging to chlorophyta (green), phaeophyta (brown) and rhodophyta (red) groups has been discussed in the following sections.
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2.3.1. Biosorption of uranium by green algae Green algae are a diverse group of photosynthetic microorganisms ranging from unicellular flagellates to complex multicellular forms. Green algae belonging to the genera Chlorella, Spirulina and Nostoc have been effective in the removal of uranium ions from aqueous solutions. Live and dead cells of Chlorella vulgaris brought about rapid biosorption of uranium between pH 3.0 to 6.0 in mineral medium with low phosphate concentration (Vogel et
ACCEPTED MANUSCRIPT al., 2010). However, after prolonged incubation (48 to 96 h), live cells caused the mobilization of bound uranium by releasing organic acids. C. vulgaris did not adsorb uranium from natural or artificial sea water at pH 8.0. The observed inhibition of uranyl uptake was due to the formation of stable uranyl-carbonate complexes, [UO 2 (CO 3 )2 2-] and [UO 2 (CO 3 )3 4-] (Nakajima et al., 1979).
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In contrast, uranyl adsorption by C. vulgaris immobilized in polyacrylamide gels was unaffected
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by pH. The immobilized cells could recover 100% U from sea water spiked with 10 mg l-1 U
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(Nakajima et al., 1979). In yet another investigation, uranium (UO 2 2+ ions) solubilized from
and Nostoc linckia (Cecal et al., 2012).
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2.3.2. Biosorption of uranium by brown algae
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uranium ores and sludges were found to be sequestered within the cells of Spirulina platensis
Brown algae are multicellular in nature and are prevalent in marine ecosystems. The
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brown color displayed by this group of algae is due to the presence of fucoxanthin (Groisillier et
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al., 2014). The biomass of Sargassum fluitans could adsorb up to 105 mg g-1 of uranium in 30 days under continuous flow through conditions (Yang and Volesky, 1999). The biomass of
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Padina sp. and Padina pavonia has been demonstrated for uranyl sorption (Khani, 2011; Aytas
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et al., 2014). Cystoseria indica has also demonstrated good performance for uranyl biosorption under acidic conditions from pH 2.5 to 4.0 (Table 2) (Khani et al., 2006; Khani et al., 2008;
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Ghasemi et al., 2011; Gök et al., 2017). Chemically modified biomass of Cystoseira sp. did not affect the process of uranium adsorption from simulated wastewater samples containing different metal ions. The biosorption process was exothermic and spontaneous as indicated by negative values of enthalpy and free energy changes (Gök et al., 2017). Alginate, a polysaccharide, derived from some brown algae has also been used for biosorption of U from aqueous solutions (Gok and Aytas 2009; Yu et al., 2017).
ACCEPTED MANUSCRIPT 2.3.3. Biosorption of uranium by red algae Natural occurrence of red algae on sea coasts and estuarine waters in large quantities allows this biomass to be used as an inexpensive biosorbent. The red alga Catenella repens showed encouraging results for biosorption of uranium under acidic conditions displaying
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maximum loading capacity of 303 mg U g-1 from solution containing 100 mg l-1 U (Bhat et al.,
ores and sludges (Cecal et al., 2012). 3. Bioprecipitation of uranium offers
a
viable
remediation
technique
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Bioprecipitation
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2008). Porphyridium cruentum, another red alga, has been shown to remove UO 2 2+ ions from
for
controlling
metal
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contamination. Microbial bioprecipitation of metals is often an enzymatic process. Phosphatases produced by microorganisms hydrolyze organic phosphate substrates and release inorganic
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phosphate (Pi). The latter interacts with metals in solutions and precipitates them in the form of
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insoluble minerals, generally metal phosphates (Lunt et al., 2007). Therefore, bioprecipitation is often referred as biomineralization as the end product is an insoluble mineral.
Alternatively, the
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degradation of polyphosphate (phosphate polymer) resulting in phosphate release and metal
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bioprecipitation has also been reported in microbes (Renninger et al. 2004; Acharya et al. 2017). Several environmental bacteria isolated from uranium enriched or contaminated sites have been
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shown to precipitate uranium as uranyl phosphate owing to their constitutive phosphatase activity (Martinez et al., 2007; Beazley et al., 2007; Merroun et al., 2011; Newsome et al., 2015). The crystalline uranium precipitates resulting from bioprecipitation of U show no alteration in the U(VI) redox state in contrast to reductive precipitation by dissimilatory metal reducing bacteria under anaerobic conditions (Wall and Krumholz, 2006). U immobilization as U(VI) phosphate (autunite) minerals under aerobic conditions can be utilized for alleviating uranium at
ACCEPTED MANUSCRIPT contaminated sites for extended periods over a wide pH range as they are not susceptible to changes in oxidation conditions. The following section highlights the potential of different microbial groups for bioprecipitation of uranium (Fig. 1). 3.1. Bioprecipitation of uranium by bacteria
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Bacteria, owing to their diversity and distinct metabolic properties, are capable of
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modulating biogeochemistry of metals. A plethora of bacteria from different geographic locations
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can precipitate aqueous uranium in the form of insoluble uranium phosphates as detailed in the following part of the review.
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3.1.1. Bioprecipitation of uranium by Gram positive bacteria
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Gram positive bacteria belonging to the genus Bacillus isolated from radionuclide and metal polluted terrestrial locations have demonstrated uranium precipitation via constitutive acid
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phosphatase activity (Beazley et al., 2007; Martinez et al., 2007). These studies involved the use
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glycerol-3-phosphate (G3P) as a substrate for enhancing phosphatase activity (Fig. 1). Bioprecipitation of U(VI) in Bacillus sphaericus JG-7B (isolated from uranium mining waste
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sediment) at pH 3.0 and 4.5 occurred via acid phosphatases in the absence of organic phosphate
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substrates. Phosphorylated biomolecules such as nucleic acids that were released from dead cells were proposed to serve as substrates for the phosphatases (Merroun et al., 2011). In most of the
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cases, uranium was precipitated in the form of stable and insoluble autunite/meta-autunite mineral phases (Merroun et al., 2006; Beazley et al., 2007; Martinez et al., 2007; Merroun et al., 2011; Theodorakopoulos et al., 2015). A facultative anaerobic bacterium Paenibacillus sp. JGTB8 isolated from uranium mining waste pile mineralized U under aerobic conditions (Reitz et al., 2014). Under anaerobic settings, mineral formation in JG-TB8 was not observed at pH 4.5 or 6.0 due to significant suppression of the constitutive phosphatase activity.
ACCEPTED MANUSCRIPT Microbacterium oxydans isolated from Siberian radioactive waste depository precipitated uranium extracellularly at pH 4.5 whereas uranium was bound to organophosphate groups such as fructose 6 phosphates at pH 2.0 (Nedelkova et al., 2007). A Microbacterium strain isolated from Chernobyl soil sample exhibited high tolerance to U and demonstrated intracellular
mechanism (Theodorakopoulos et al., 2015).
are
several
Alphaproteobacteria,
reports
on
various
Gammaproteobacteria
Gram
and
negative
bacteria
Deltaproteobacteria
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There
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3.1.2. Bioprecipitation of uranium by Gram negative bacteria
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accumulation in the form of autunite mineral formed as a result of phosphate released by active
belonging
classes
to
exhibiting
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bioprecipitation of uranium which are detailed in this section.
Two Alphaproteobacteria namely, Sphingomonas sp. S15-S1and Caulobacter crescentus
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exhibited uranium precipitation by the virtue of constitutive acid and alkaline phosphatase
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activities, respectively (Merroun et al., 2006; Yung and Jiao, 2014). phoY gene encoding the periplasmic enzyme, PhoY (responsible for whole cell phosphatase activity) identified in the
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annotated genome of C. crescentus facilitated uranyl precipitation. A PhoY deletion mutant of C.
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crescentus neither released Pi in presence of glycerol-2-phosphate (G2P) nor precipitated U. The results established the role of PhoY in uranium precipitation (Yung and Jiao, 2014). (reclassified
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Citrobacter
later
as
Serratia,
Pattanapipitpaisal et
al.,
2002)
a
Gammaproteobacterium in the presence of glycerol phosphate was able to bioprecipitate uranium (Macaskie et al., 1992; Macaskie et al., 1994). Uranyl phosphate mineral phases are considered to be very stable and are not amenable to oxidative remobilization unlike the products of uranium reduction which can be re-oxidized to mobile U (Senko et al., 2002). Detailed analysis of uranium precipitation in Citrobacter revealed that monophosphate groups harbored
ACCEPTED MANUSCRIPT within the lipopolysaccharide (LPS) layer facilitated the initial nucleation. Further consolidation and growth of crystalline uranyl phosphate mineral was achieved by enzymatically generated phosphate groups (Macaskie et al., 2000). Expression of higher levels of phosphatases alone did not result in uranyl precipitation (Basnakova et al., 1998). Rather, a combination of cell surface
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architecture and phosphatase overproduction facilitated efficient U precipitation. Whole cells of
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Citrobacter immobilized in polyacrylamide matrices demonstrated impressive U loading of 9 g
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U g-1 of dry weight in the form of uranyl phosphate as a result of phosphatase mediated U precipitation (Macaskie et al., 1991). Most of the studies on uranium bioprecipitation have been
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demonstrated under aerobic conditions. In contrast, a study involving Serratia sp. isolated from
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sediments of a nuclear site exhibited uranyl precipitation under anaerobic conditions (Newsome et al., 2015). Constitutive phosphatase activity and uranium bioprecipitation in yet another
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Serratia strain were uncompromised even after exposure to high levels of ionizing radiation
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presenting its utility for nuclear waste decontamination (Paterson‐Beedle et al., 2012). Similarly, other members of Gammaproteobacteria such as Acidithiobacillus ferrooxidans, Rahnella sp.,
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Pseudomonas aeruginosa, Rhodonobacter, Acinetobacter sp. and Pseudomonas fluorescens and
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Deltaproteobacteria member like Myxococcus xanthus have demonstrated uranyl precipitation via Pi derived through phosphatase activity or polyphosphate hydrolysis (Fig. 1) (Merroun et al.,
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2003; Renninger et al., 2004; Jroundi et al., 2007; Beazley et al., 2007; Choudhary and Sar, 2011; Sousa et al., 2013; Sowmya et al., 2014; Krawczyk-Bärsch et al., 2015). In addition to the reports on uranium bioprecipitation by planktonic cells, biofilms of P. fluorescens precipitated U in the form of autunite (Krawczyk-Bärsch et al., 2015). A series of interesting responses with respect to prolonged uranium exposure in the filamentous cyanobacterium, A. torulosa, was observed leading to uranium precipitation and
ACCEPTED MANUSCRIPT biomass regeneration (Fig. 3) (Acharya et al., 2017). Cells of A. torulosa accumulated uranium in polyphosphate bodies within 24 h of U (100 µM) exposure under phosphate limited conditions. Further incubation of the cells (48 h) caused cell lysis and hydrolysis of polyphosphates with concomitant release of U and Pi. Cell bleaching, akinete formation followed
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by uranium precipitation were observed following 120 h of uranyl exposure in this organism.
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The precipitated U (in the form of autunite mineral) settled at the bottom of the reaction vessel
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and provided a U “free” environment. Subsequently, inducible alkaline phosphatase activity in the akinetes led to germination of akinetes and complete regeneration of A. torulosa culture by
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384 h (16 d) of U exposure. This phenomenon presented unusual insights into various metabolic
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responses adopted by A. torulosa to resist and survive under sustained uranium toxicity (Acharya et al., 2017).
Such responses can be utilized to mitigate uranium contamination for prolonged
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period.
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3.2. Bioprecipitation of uranium by yeast and fungal species Bioprecipitation of uranium has been observed in several yeasts and fungal species. In S. formation
of HUO 2 PO 4 ·4H2 O
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cerevisiae,
(H-autunite) was observed
because of the
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complexation of adsorbed uranium with the phosphate released from the cells. The mineral was formed at the ruptured regions of cell surfaces due to the localized saturation of the released
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phosphate (Ohnuki et al., 2005). Several other yeast species such as Kluyveromyces lactis, Pichia acacia, Cryptococcus podzolicus, Cryptococcus filicatus, Candida sake and Candida argentea have also been studied with respect to uranium bioprecipitation (Liang et al., 2016). Uranium phosphate precipitates were observed on the yeast cell surface on addition of organic phosphate substrates such as G2P or phytic acid in the medium. The precipitated uranium complexes were identified
as
chernikovite,
meta-ankoleite
K 2 (UO 2 )2 (PO 4 )2 ·6(H2 O),
bassetite
ACCEPTED MANUSCRIPT Fe++(UO 2 )2 (PO 4 )2 ·8(H2 O) and uramphite (NH4 )(UO 2 )(PO 4 )·3(H2 O) (Liang et al., 2016). Filamentous fungi like Aspergillus niger and Paecilomyces javanicus precipitated uranium significantly despite of their reduced growth in uranium containing media. The hyphal matrix of fungi served as the site for localization of uranyl phosphate minerals (Liang et al., 2015).
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4. Bioreduction of uranium
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Microbial reduction of metals involves the immobilization of potentially toxic soluble
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metals and radionuclides in their insoluble form by alteration of their oxidation states. Early reports on uranium reduction focused on abiological reactions with reductants such as sulfides,
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molecular hydrogen (H2 ) and organic compounds (Maynard, 1983). However, it is now evident
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that some bacteria can mediate reduction of the soluble oxidized form of uranium U(VI) to the insoluble tetravalent form, U(IV) (Wall and Krumholz, 2006; Newsome et al., 2014).
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Bioreduction is mostly observed in anaerobic bacteria which grow using uranium as potential
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electron acceptor or in other words, these microbes are said to respire uranium (Majumder and Wall, 2017). Since uranium bioreduction results in the formation of insoluble mineral uraninite,
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UO2 , it is also interchangeably termed as biomineralization. Uranium reduction in various
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microbes is detailed in this section (Fig. 1). 4.1. Bioreduction of uranium by Gram positive obligate anaerobic bacteria
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Several species of Clostridia have the inherent property of reducing U(VI) to U(IV) under acidic conditions (pH 5-6) (Gao and Francis, 2008). Clostridium is proposed to be one of the most important organisms for effectuating uranium reduction under natural conditions such as uranium pit water (Petrie et al., 2003), military facilities (Dong et al., 2006) and nuclear material storage sites (Madden et al., 2007). Fermentative processes in Clostridia are associated with U reduction (Petrie et al., 2003).
ACCEPTED MANUSCRIPT A spore forming sulfate-reducing bacterium, Desulfotomaculum reducens MI-1 isolated from metal contaminated sediments could grow by using U(VI) as the terminal electron acceptor in the absence of sulfate (Tebo and Obraztsova,1998). This organism could couple the oxidation of organic compounds with reduction of U(VI) to U(IV) and other metals such as manganese
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[Mn(IV) to Mn(II)], iron [Fe(III) to Fe(II)] or chromium [Cr(VI) to Cr(III)] for its growth.
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Spores of this bacterium could also bring about bioreduction of uranium by using H2 as the
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electron donor (Junier et al., 2009).
4.2. Bioreduction of uranium by Gram negative facultative anaerobic bacteria
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Mechanisms for uranium reduction in bacteria are not fully understood. Membrane
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associated electron transport respiratory (ETR) systems containing c-type cytochromes play an important role in uranium reduction (Lloyd, 2003). U reducing facultative anaerobic bacteria
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prevents the formation of sulfides and hydrogen sulfide complexes that have toxic and inhibitory
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effects on cellular metabolism (Mtimunye and Chirwa, 2014). In such microorganisms, extracellular U(VI) reduction is observed which involves electron flow through NADH-
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dehydrogenase, a primary electron donor associated with the ETR system. Salmonella
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subterranean (an acid tolerant bacterium and an important component of U(VI) reducing enrichment culture isolated from nitrate and uranium contaminated subsurface sediments)
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demonstrated significant reduction of uranium at pH 4.5 (Shelobolina et al., 2004). Members of the genus Shewanella have been extensively studied with respect to bioreduction of uranium. Initial studies on Shewanella putrefaciens demonstrated that the bioreduction of uranium was facilitated by enzymatic reactions coupled to electron transport chains within the organism (Gorby and Lovley, 1992). Shewanella alga, an iron reducing bacterium brought about reduction of uranium when complexed with multidentate aliphatic
ACCEPTED MANUSCRIPT ligands (malonate, oxalate and citrate) rather than monodentate acetate groups (Ganesh et al., 1997). This bacterium was unable to reduce uranium complexed with aromatic ligand and metal chelator such as tiron (4,5-dihydroxy-1,3-benzene disulfonic acid). Another important factor affecting uranium bioreduction kinetics is the bioavailability of uranium to the bacteria. U(VI)
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reduction by Shewanella oneidensis MR-1 in presence of sodium bicarbonate (NaHCO 3 ) showed
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slower reduction kinetics. The adsorption of U onto bacteria at pH 7.0 decreased with an increase
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in the concentration of NaHCO 3 due to the formation of soluble, negatively charged uranyl carbonate species (Sheng et al., 2011). A series of biological and chemical redox transformations
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in S. oneidensis including absorption of U(VI) to cell surface, formation of U(VI) nanowires
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[meta-schoepite (UO 3 .2H2 O)] and subsequent reduction to U(IV) nanoparticles [uraninite (UO 2 )] presented a novel strategy for U bioremediation in contaminated aquifers (Jiang et al., 2011).
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The two step uranium reduction was superior to the previously reported one step reduction
ED
process (Lovley et al., 1991). Mixed bacterial cultures derived from natural sediments and pure cultures of Desulfosporosinus species, Geobacter sulfurreducens, S. oneidensis MR-1 and S.
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putrefaciens CN32 are also reported to form biogenic uraninite nanoparticles (Suzuki et al.,
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2002; Renshaw et al., 2005; Senko et al., 2007; Bargar et al., 2008). In S. oneidensis MR-1, ctype cytochromes played an important role in reducing U(VI) to extracellular UO 2 nanoparticles
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(Marshall et al., 2006). The outer membrane (OM) associated decaheme cytochrome MtrC was essential for transferring electrons to U(VI) thereby reducing it to U(IV). Mutants of S. oneidensis MR-1 lacking mtrC (coding for cytochrome MtrC) or omcA (encoding for OM decaheme c-type cytochromes) displayed slower uranium reduction compared to the wild type strains (Marshall et al., 2006). Similarly, endecaheme c-type cytochrome (UndAHRCR-6) present in the OM of Shewanella sp. were involved in the extracellular reduction of iron [Fe(III)]
ACCEPTED MANUSCRIPT oxides and uranium [U(VI)] (Shi et al., 2011). Biofilms of Shewanella sp. HRCR-1 were also able to reduce uranium efficiently with the help of outer membrane c-type cytochromes present in the extracellular polymeric substances (Cao et al., 2011). However, S. oneidensis MR-1 and its biofilm exhibited limitations in reduction of U(VI) in flow and batch reactors (Sani et al., 2008).
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Apart from genetic manipulations, alternative organic materials/modules have been shown to
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facilitate and enhance uranium bioreduction rates. Anthraquinone-2,6-disulfonate provided as
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humus substitute improved the uranium reduction rates in S. oneidensis (Liu et al., 2015).
reduction rate of U(VI) (Suzuki et al., 2010).
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Similarly, FMN (Flavin mononucleotide) secreted by Shewanella species accelerated the
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4.3. Bioreduction of uranium by Gram negative obligate anaerobic bacteria In this section, the potential of Gram negative bacteria Geobacter metallireducens
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(previously designated as strain GS-15) and Anaeromyxobacter dehalogenans along with variety
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of sulfate reducing bacteria (SRB) in bringing about uranium bioreduction has been discussed. Geobacter species can mediate coupled reduction of Fe(III) and U(VI) in anaerobic subsurface
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environments. G. metallireducens, a Fe(III) reducing bacterium, showed direct reduction of
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U(VI) in contrast to indirect reduction of Fe(III). The reduction process promoted bacterial growth and the culture ceased to grow when uranium was depleted from the growth medium
AC
(Lovley et al., 1991). This organism reduced uranium in groundwater samples and the deposits of uraninite were observed on cell surfaces (Gorby and Lovely, 1992). Nitrate, a co-contaminant associated with uranium has a deleterious effect on reduction processes (Riley and Zachara, 1992) and it was necessary to add sufficient quantities of acetate (as an electron donor) initially to reduce nitrate (Finneran et al., 2002). Reduction and precipitation of U on surfaces harboring abundant electron donors cause damage to the cell envelope and interferes with the associated
ACCEPTED MANUSCRIPT functions. In order to safeguard cellular integrity and viability, conductive pili in Geobacter sp. demonstrated extracellular reduction of U(VI) to U(IV) (Cologgi et al., 2011). Pili were identified as primary uranium reductases that accepted electrons from the cell envelope and ctype cytochromes (Fig. 1). These served as electrical conduits between the cells and uranium.
in turn,
reduced
cell viability and cellular respiration
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periplasmic mineralization which,
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The lack of pili (Pil A- mutant) impaired U reduction in Geobacter and caused enhancement of
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(Reguera, 2012). A sulfate reducing species of Geobacter namely, G. sulfurreducens initially reduced U(VI) to a U(V) intermediate which on disproportionation resulted in the formation of
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tetravalent uranium (Renshaw et al., 2005). Cytochromes associated with outer surfaces (c-type)
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played an important role in U reduction in this organism (Orellana et al., 2013). Moreover, periplasmic cytochrome MacA, a diheme c-type cytochrome peroxidase, was found to be
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essential for U(VI) reduction in G. sulfurreducens and a fraction of the reduced uranium was
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localized in the periplasm (Shelobolina et al., 2007). Growth yields of Geobacter lovleyi and G. sulfurreducens were lowered when U(VI) was used as electron acceptor suggesting that this
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process of reduction imposed additional burden to the growing cells (Sanford et al., 2007).
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Geobacter species have played a significant role in bioremediation of uranium under microcosm and in situ conditions. Microcosm studies using soil samples from Oak Ridge, FRC,
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TN, U.S. showed the enrichment of the members of Geobacteraceae family that could reduce uranium under low bicarbonate concentrations (1 mM) in the presence of ethanol as the electron donor (Luo et al., 2007). Uranium bioreduction continued up to 50 days during a field trial at Rifle, CO, U.S. (Anderson et al., 2003). With acetate supplementation (as an electron donor), concentrations of U(VI) were observed to decrease from 0.4 to those below 0.18 μM which was considered to be the maximum contaminant limit. In situ reduction of U(VI) in this site was not
ACCEPTED MANUSCRIPT effective at greater depths (4.0 to 5.2 m) as the substantial proportion of available uranium was not amenable for bioreduction (Ortiz-Bernad et al., 2004). The microbial community at a uranium contaminated field site in Rifle Integrated Field Research Centre (IFRC), CO, U.S. was also dominated by Geobacter species when biostimulated with acetate. These bacteria effectively
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brought about dissimilatory uranium reduction (Mouser et al., 2009).
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An anaerobic myxobacterium isolate obtained from Oak Ridge FRC, TN, U.S. was
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identified as A. dehalogenans. This culture reduced uranium by using H2 as the electron donor in contrast to acetate that was utilized by Geobacter (Wu et al., 2006). While alternative electron
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acceptors such as Fe(III) citrate or Fe(III) oxide inhibited U(VI) reduction in A. dehalogenans,
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fumarate and 2-chlorophenol did not show such inhibitory effect (Wu et al., 2006). The SRB, Desulfovibrio desulfuricans, demonstrated rapid and direct reduction of uranium in the presence
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of lactate or H2 supplied as electron donors (Lovley and Phillips, 1992a). U(VI) and sulfate
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reduction occurred simultaneously and reduced uranium was observed in the form of extracellular uraninite. D. desulfuricans cells contained within a semipermeable membrane
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showed rapid reduction of uranium at concentrations as high as 24 mM (Lovley and Phillips,
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1992b). Bicarbonate extraction of uranium from contaminated soils and subsequent microbial reduction by D. desulfuricans demonstrated a potential strategy for concentrating U from
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contaminated soil and sediments (Phillips et al., 1995). Cytochrome c3 appeared to be integral for in vivo electron pathway that was involved in the reduction of U(VI) to U(IV) in D. desulfuricans. A cytochrome c3 mutant strain of D. desulfuricans G20 was less effective (almost by 50%) in uranium reduction than wild type strains in presence of lactate or pyruvate as electron donors (Payne et al., 2002). A variety of sulfate reducing bacteria (SRB) were analyzed for their ability to reduce uranium (Lovley et al., 1993a). Among different isolates, Desulfovibrio vulgaris
ACCEPTED MANUSCRIPT was the most effective uranium reducing organism followed by Desulfovibrio baculatus, Desulfovibrio sulfodismutans and Desulfovibrio baarsii (Lovley et al., 1993a). Uranium reduction by all these SRBs was highly dependent on the supplementation of electron donors. D. vulgaris was unable to grow in presence of U(VI) as the sole electron acceptor, precipitated
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uranium as uraninite and reduced sulfate and uranium simultaneously (Lovley et al., 1993b)
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similar to D. desulfuricans (Lovley and Phillips, 1992a). Another SRB, Desulfosporosinus sp.
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(isolate P3) and its type strain Desulfosporosinus orientis (DSM 765) showed enzymatic reduction of uranium using lactate or H2 as electron donor (Suzuki et al., 2004).
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A field study related to growth and activity of indigenous microorganisms in the presence
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of acetate, glucose or ethanol was carried out at Oak Ridge, FRC, TN, U.S. (Istok et al., 2004). Supplementation with these electron donors resulted in the development of an anaerobic and
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reducing environment that favored reduction of NO 3 -, Fe(III) and U(VI). NO 3 - dependent,
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microbially mediated U(IV) oxidation is an important process in modulating the stability of bioreduced U(VI) at sites contaminated with high concentration of nitrates. In situ uranium
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bioremediation trials with acetate supplementation were carried out at Rifle, IFRC, CO, U.S.
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Rapid reduction of uranium prevalent in the form of Ca-UO 2 -CO3 ternary complexes was achieved by iron and uranyl respiring Geobacter species (Williams et al., 2011). In a later study,
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at the same site, U reduction was observed in presence of acetate, lactate, hydrogen release compound (HRC) or vegetable oil. The composition of microbial communities varied with different donors (Barlett et al., 2012). Another report at Oak Ridge IFRC, TN, U.S. describes U(VI) bioreduction over an extended (one year) period. Emulsified vegetable oil (EVO) amendments caused effective bioreduction at depths up to 50 m (Watson et al., 2013). Recent studies at Oak Ridge, TN, U.S. have demonstrated U(VI) bioreduction in iron oxide-rich
ACCEPTED MANUSCRIPT sediments after ethanol supplementation (Li et al., 2018). Major problem associated with in situ bioremediation of uranium contaminated sites is the fate of reduced U(IV) in sub-surfaces which re-oxidizes to U(VI) form and affects the process economy negatively. A recent review focuses on the challenges and complexities associated with remediation of uranium contaminated soils
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(Selvakumar et al., 2018). Factors such as soil type, presence of organic matter, co-ions along
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with soil pH and uranium species play a major role in determining the efficiency of
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bioremediation methods and should be considered while designing strategies for uranium removal from contaminated soil.
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5. Uranium bioremediation by genetically engineered bacteria
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The use of genetically engineered microorganisms (GEMs) for bioremediation of metal polluted environments has become important in the recent years. There are several reports
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highlighting the potential of GEMs for removing toxic metals and radionuclides (Garbisu and
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Alkorta, 1999; Urgun-Demirtas et al., 2006; Menn et al., 2008; Kulkarni et al., 2013). Relevant genes have been effectively expressed in appropriate hosts and the resulting GEMs have
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demonstrated remarkable bioremediation efficiencies (Appukuttan et al., 2006, 2011; Nilgiriwala
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et al., 2008; Kulkarni et al., 2013, 2016). Majority of the studies utilizing genetically engineered microorganisms have focused on
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phosphatase mediated uranium precipitation rather than on biosorption or bioreduction. For example, Escherichia coli cloned with a non-specific acid phosphates gene, phoN showed superior (2.5 times) uranium precipitation than Citrobacter sp. constitutively expressing the phoN gene (Basnakova et al., 1998). The uranyl precipitates in the recombinant E. coli cultures were identified as HUO 2 PO4 . The radiosensitivity of E. coli, Citrobacter sp. or Pseudomonas sp. seemed to limit their usage in treatment of radioactive waste. Recombinant bacteria expressing
ACCEPTED MANUSCRIPT phosphatases and capable of precipitating uranium under acid or alkaline conditions even after being exposed to high doses of ionizing radiation have been reported (Appukuttan et al., 2006; Appukuttan et al., 2011; Kulkarni et al., 2013). Deinococcus radiodurans is known to tolerate extremely high doses of ionizing radiations well above those encountered under natural
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conditions (Battista, 1997). Recombinant cells of D. radiodurans expressing a non-specific acid
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phosphatase, PhoN (DrPhoN) were shown to precipitate uranium efficiently in presence of G2P
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under acidic conditions (Appukuttan et al., 2006). However, recombinant cells of D. radiodurans exhibited slower uranium precipitation (70% after 26 h) as compared to E. coli cells expressing
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PhoN (70% in 2 h) when exposed to aqueous solutions containing 0.8 mM uranyl nitrate at pH
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5.0 (Appukuttan et al., 2006). The multilayered cell envelope of D. radiodurans apparently decreased the accessibility of the organophosphate substrate to the periplasmic PhoN as
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compared to E. coli. Recombinant D. radiodurans cells retained 90% of their uranium 60
Co gamma rays
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precipitation abilities even after exposure to very high doses (6 kGy) of
presenting their potential for treating radioactive nuclear waste. Radiation sensitive E. coli
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clones, on the other hand, lost their ability to precipitate U following exposure to lower doses (1
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kGy) of gamma rays (Appukuttan et al., 2006). Moreover, the lyophilized DrPhoN cells showed appreciable loading of bioprecipitated uranium up to 5.7 g U g-1 dry weight. This property was
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retained even after 6 months of storage at room temperature. Fresh (non-lyophilized) cells on the other hand, lost their ability to precipitate U after 8-10 days of storage at 4˚C (Appukuttan et al., 2011) establishing the enhancement of shelf-life of lyophilized cells for application purposes. It was observed that the recombinant D. radiodurans cells expressing PhoN from a radiation inducible ‘ssb’ promoter exhibited enhanced phosphatase activity following exposure to gamma
ACCEPTED MANUSCRIPT radiation and removed uranium more rapidly as compared to the cells expressing PhoN from a radiation non-inducible Deinococcal groESL promoter (Misra et al., 2014). Uranium bioprecipitation from alkaline solutions has also been attempted by using recombinant bacteria. Recombinant E. coli or D. radiodurans cells expressing Sphingomonas
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PhoK (alkaline phosphatase) demonstrated high enzyme activity and precipitated U from
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alkaline solutions (pH 9.0) very effectively (Nilgiriwala et al., 2008; Kulkarni et al., 2013;
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Kulkarni et al., 2016). Recombinant D. radiodurans strain expressing PhoK revealed very high loading (10.7 g U g-1 dry weight of cells) of uranium (Kulkarni et al., 2013). These cells 60
Co gamma rays. When
in calcium alginate, they exhibited superior uranium removal from alkaline
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immobilized
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precipitated uranium effectively even after exposure to 6 kGy dose of
solutions (Kulkarni et al., 2013).
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There are few reports on development of recombinant bacterial strains for uranium Asn
128T yr change in the chromate
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bioreduction and bioaccumulation. A single amino acid
reducing enzyme ChrR6 from E. coli resulted in significant improvement in the kinetics of
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uranyl reduction and minimized the chances of reoxidation and resolubilization of the reduced
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product (Barak et al., 2006).
Metallothioneins (MT) are small, cysteine-rich, intracellular, metal binding proteins that
E.
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are implicated in metal detoxification (Pettersson et al., 1988; Daniels et al., 1998). Recombinant coli cells expressing cyanobacterial (Synechococcus) metallothionein, SmtA exhibited
increased
tolerance towards uranium (Acharya and
Blindauer, 2016). However, SmtA
homologue, namely NmtA from another cyanobacterium, Anabaena sp. strain PCC 7120 when expressed in E.coli did not confer any notable uranium tolerance (Divya et al., 2018). The recombinant SmtA was able to bind uranium in the presence of excess carbonate ions under in
ACCEPTED MANUSCRIPT vitro conditions via its glutamate and aspartate residues without compromising its secondary or tertiary structure (Acharya and Blindauer, 2016). The uranyl binding affinity of recombinant SmtA, however, was lesser (KD = 100 pM) than that of a protein engineered for high uranyl binding, namely, super uranyl binding protein (SUP) (KD = 7.4 fM). SUP was obtained by
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multiple mutations of the ligand residues (to aspartate/asparagine or glutamate/glutamine) in a
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small protein of unknown function in Methanobacterium thermoautotrophicus (Zhou et al.,
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2014). SUP when fused with maltose binding protein could remove 90% of U from synthetic sea water in presence of all major ions. SUP was displayed on E. coli surface with OmpA (outer
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membrane protein) as a fusion protein and the resulting recombinant E. coli cells could sequester
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60% of U from synthetic sea water (Zhou et al., 2014). Similarly, a uranyl responsive DNA binding protein was designed by engineering the nickel (IIs) responsive protein, NikR from E.
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coli. The resulting recombinant NikR protein bound uranyl with a dissociation constant K D= 53
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nM and selectively bound to DNA in presence of uranium (Wegner et al., 2009). Although genetically engineered microorganisms display potential interactions with U,
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their exploitation for bioremediation applications requires further validation.
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6. Understanding uranium interactions in microbes via ‘omics’ based approaches Recent advances in analytical and bioinformatics techniques have made it possible to
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understand gene/protein expression under different stress conditions via genomic and proteomic approaches. Genomics and metagenomics are important in establishing genetic variability, population structure and ecological functions of microorganisms (Riesenfeld et al., 2004; Xu, 2006). We discuss here, genomics, transcriptomics and proteomics studies in microbes in response to the uranium exposure (Fig. 1). 6.1. Genomics, transcriptomics and metagenomics investigations in the presence of uranium
ACCEPTED MANUSCRIPT Such studies allow assessment of the impact of metals on microbial communities (Akob et al., 2006; Suriya et al., 2017). Microorganisms living in contaminated sites exhibit different strategies for counteracting toxic effects of metals. For example, the occurrence of P IB-type ATPases (metal transporters) was established in uranium tolerant, aerobic heterotrophic bacteria
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isolated from unmined uranium rich deposits in North-East India as a result of horizontal gene
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transfer (Nongkhlaw et al., 2012).
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Gene expression studies have helped in understanding the mechanism of uranium tolerance in microorganisms. Transcriptome profiling of Geobacter uraniireducens and S.
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oneidensis MR-1 has illustrated a varied stress response in the presence of uranium (Bencheikh-
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Latmani et al., 2005; Holmes et al., 2009). Transposon mutagenesis in C. crescentus revealed the importance of outer membrane transporters (rsaFa and rsaFb ), a stress-responsive transcription
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factor (cztR), or a ppGpp synthetase/hydrolase (spoT) in conferring uranium tolerance. RsaF a and
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RsaFb were shown to efflux U and maintain the membrane integrity during U stress (Yung et al., 2015). Transcriptomics studies on Metallosphaera prunae (a thermoacidophilic bacterium
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isolated from a uranium mine in Thuringen, Germany) exposed to high concentrations of soluble
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uranium (1,238 mg l-1 ) were carried out (Mukherjee et al., 2012). Substantial degradation of cellular RNA and abortion of transcriptional and translational processes were observed in the
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organism within 15 min of U exposure possibly for resisting uranium toxicity. However, RNA integrity was restored within 60 min of uranium treatment. Possible role of siderophores in U sequestration in M. prunae was suggested as the genes related to iron complex transport system were found to be significantly induced during U exposure (Mukherjee et al., 2012). Metagenomic analysis of microbial community from uranium contaminated sub surfaces of Oak Ridge FRC, TN, U.S., Ervas Tenras Pinhel Guarda, Portugal, Cauvery River, Tamil
ACCEPTED MANUSCRIPT Nadu,
India
and
uranium ore
deposit
of Domiasiat,
India,
showed
abundance
of
Gammaproteobacteria (Akob et al., 2006; Caetano et al. 2014; Suriya et al., 2017; Kumar et al., 2013b). Similar observations were made with respect to uranium polluted sites at Bessines, Limousin, France (Mondani et al., 2011) and uranium mines at Jaduguda, Bagjata and Turamdih
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located in East Singhbhum district, Jharkhand, India (Dhal et al., 2011). Proteobacteria like A.
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dehalogenans and Geobacter were found in abundance at Oak Ridge National Laboratory, TN,
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U.S. (North et al., 2004) and the dominance of Geobacter species like Geobacter bemidjiensis, G. metallireducens, G. sulfurreducens PCA, G. lovleyi SZ, G. uraniireducens Rf4, Geobacter
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strain FRC-32 and Geobacter strain M21 was observed at U contaminated sites in Rifle, CO,
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U.S. (Wilkins et al., 2009). Members of Rhodocyclaceae family of Proteobacteria group were found in samples collected from different locations of U contaminated sites of Portugal (Martins
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et al., 2010). Burkholderia sp. strain SRS-W-2-2016 (Proteobacteria group) isolated from
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Savannah River Site, U.S. displayed ecologically relevant genomic traits such as substrate binding proteins, permeases, transport regulators and efflux pumps to restrict metals in the
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extracellular environment or prevent their cellular uptake. This aided survival of the bacterium in
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the presence of mixed contaminants (Pathak et al., 2017). Apart from Proteobacteria, Actinobacteria and Firmicutes have also been reported from
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Oak Ridge, FRC, TN, U.S. and uranium mine in Jaduguda, India (Bollmann et al., 2010; Islam and Sar, 2011). Similarly, bacterial community of soil samples collected from T22 trench of the Chernobyl, Ukraine, was dominated by Firmicutes (Chapon et al., 2012). Desulfotomaculum like microorganisms (Firmicutes) were dominant in uranium mill tailings disposal site at Shiprock, New Mexico (Chang et al., 2001). Proteobacteria and Acidobacteria were reported from soil samples of uranium ore deposit of Domiasiat in Northeast India (Kumar et al., 2013b).
ACCEPTED MANUSCRIPT Metagenomic analysis of a radioactive contaminated soil at uranium tailings from southern China showed structural and functional diversity of the soil bacterial community with abundance of Actinobacteria and Proteobacteria groups and proteins related to membrane transport and carbohydrate metabolism (Yan et al., 2016).
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A recent study on microbial community analysis of uranium contaminated aquifers at
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Oak ridge IFRC, TN, U.S. has shown the dominance of SRB such as Desulfococcus,
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Desulfobacterium and Desulfovibrio (Zhang et al., 2017). The functional diversity of microorganisms was shown to decrease in the groundwater samples at higher uranium
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concentrations (He et al., 2018). The levels of uranium contamination and ecosystem functioning
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could be predicted on the basis of such variations in microbial diversity. The effect of varying uranium concentrations (0-4000 mg kg-1 ) on microbial diversity was examined in Gulungul
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billabong near Ranger uranium mines, Australia (Sutcliffe et al., 2017a). Archaeal and bacterial
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components prevalent in methanogenic consortia were observed at higher concentrations. Members belonging to Orders Rhizobiales and Acidobacterial were inhibited by uranium. The
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genes associated with methanogenesis, nitrogen fixation and metal efflux showed a positive
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correlation with uranium concentration while the genes involved in aerobic respiration and amino acid transporters were negatively affected (Sutcliffe et al., 2017a). Another study on
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microbial communities associated
with sediments of the same site with high uranium
concentrations (4000 mg kg-1 ) showed the predominance of five bacterial genera namely, Geobacter, Geothrix, Dyella, Bacteroidetes and Chloroflexi (Sutcliffe et al., 2017b). Overall, it is evident from the cited studies that members of Phylum Proteobacteria are most dominant members in uranium contaminated environments. 6.2. Proteomic responses in presence of uranium
ACCEPTED MANUSCRIPT Proteomic studies have been carried out to understand the mechanisms involved in uranium tolerance and detoxification. In this section, comparative proteomic analysis in response to uranium observed in different groups of microorganisms has been discussed. The major alterations
in
protein
profiles
following
uranium
exposure
in
some
representative
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6.2.1. Proteomic responses of bacteria in presence of uranium
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microorganisms are highlighted in Table 3.
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Proteome analysis of C. crescentus exposed to U revealed down regulation of cell cycle regulators, motility, chemotaxis proteins and up regulation of a putative phytase (Yung et al.,
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2014). Phytase provided the requisite phosphate groups for uranium precipitation and favored
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cell survival of C. crescentus in presence of U (Yung and Jiao, 2014). Increased expression of proteins involved in general stress response or in detoxification of reactive oxygen species was
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observed in U exposed cells of A. ferrooxidans (Dekker et al., 2016). Similarly, E. coli cells
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exposed to uranyl ions showed differential accumulation of oxidative stress proteins and other proteins like NADH/quinone oxidoreductase WrbA (Khemiri et al., 2014). Low molecular
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weight thiols played an important role in protecting cells of G. sulfurreducens, E. coli and A.
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ferrooxidans against U induced oxidative stress (Orellana et al., 2014; Khemiri et al., 2014; Dekker et al., 2016). G. sulfurreducens displayed increased levels of several efflux pumps
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belonging to the RND (resistance-nodulation-cell division) family, superoxide dismutase and superoxide reductase when exposed to U (Orellana et al., 2014). Similarly, the higher tolerance of uranium in Anabaena strain L-31(LD50 of 200 µM) as compared to Anabaena strain PCC 7120 (LD50 of 75 µM) was attributed to the increased levels of oxidative stress defense protein like Manganese SOD (MnSOD) in L-31 as compared to PCC 7120 (Panda et al., 2017). Proteins involved in phosphate and iron metabolism were found to be abundant in a uranium tolerant
ACCEPTED MANUSCRIPT bacterium, Microbacterium oleivorans A9, isolated from a sample near Chernobyl nuclear power plant. The role of siderophores was proposed in uranium transportation in this bacterium (Table 3) (Gallois et al., 2018). 6.2.2. Uranium induced proteomic responses in yeasts U(VI) or
U(VI) and increased expression of specific uncharacterized proteins following U exposure
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233
238
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S. cerevisiae strain BY4743 showed reduced growth when exposed to
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(Sakamoto et al., 2007). Deletion of the phosphate transporter genes of PHO86, PHO84, PHO2, and PHO87 resulted in the uranium-sensitive (mutant) yeast strains suggesting that the phosphate
US
transporting genes contributed to uranium tolerance. Deletion of genes encoding for cell
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membrane proteins, phospholipid-binding proteins and cell wall proteins in some of the uncharacterized yeast strains demonstrated lower uranium accumulation in the mutant strains
M
indicating the positive role of the cell surface proteins in the accumulation process (Sakamoto et
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al., 2012).
6.2.3. Metaproteomics analysis of uranium exposed communities
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Metaproteomics studies of environmental samples have become possible with recent
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advances in development of protein extraction methods from groundwater, soils and sediments (Wrighton et al., 2012; Handley et al., 2013). The metaproteome profiles revealed changes in the
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microbial activities and identified active biochemical pathways and functions specific to microbial groups after biostimulation withemulsified
vegetable oil (EVO) at a uranium
contaminated site (Chourey et al., 2013) Members of Betaproteobacteria (Dechloromonas, Ralstonia, Rhodoferax, Polaromonas, Delftia and Chromobacterium) and Firmicutes and the proteins of P. putida, Geobacter species and Dechloromonas like microbes were found to be abundant in contaminated sites. Moreover, proteins responsible for ammonium assimilation,
ACCEPTED MANUSCRIPT EVO
degradation
and
polyhydroxybutyrate granule formation were prominent following
biostimulation (Chourey et al., 2013). A proteogenomic study during in situ U bioreduction by Geobacter members at Oak Ridge FRC, TN, U.S. revealed that the proteome of these strains were dominated by enzymes which converted acetate to acetyl-CoA and pyruvate throughout the
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biostimulation process (Wilkins et al., 2011). The identification of functional genes and proteins
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implicated in metal responses are fundamental in deciphering underlying molecular mechanisms
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and for developing in situ bioremediation strategies. 7. Conclusions
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In the recent times, uranium contamination resulting from anthropogenic activities has
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affected the microbes in various ways. Due to lack of clarity on the biological relevance of uranium, microbial interactions with uranium have been explored less as compared to other
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potentially toxic metals including arsenic, cadmium, mercury, silver and chromium. It is
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therefore important to understand microbial machinery in response to U to develop in situ remediation measures for controlling uranium contamination. A summary of the interactions of
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microorganisms with uranium along with ‘omics’ tools used for understanding these interactions
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are depicted in Fig. 1. The use of microbial systems is advantageous for providing cost effective solutions without the generation ofhazardous secondary wastes. Scale-up procedures and field
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trials have been undertaken on many occasions for uranium bioremediation. Recent advances in genomics, transcriptomics, metagenomics, proteomics and metaproteomics have helped us in understanding the significance of genes,
proteins or microbial communities capable of
interacting with uranium. In the future, the natural U tolerant environmental microbes as well as the microbial metabolic products obtained using ‘omics’ based approaches could assist in natural attenuation of U contamination.
ACCEPTED MANUSCRIPT Conflicts of interest The authors declare that they have no competing interests . Acknowledgments SZ and NK thank University Grants Commission for funds under University for Potential
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Excellence (UPE) Phase II. NK thanks SPPU-BARC collaborative Ph.D. Programme for
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financial support. We gratefully acknowledge the help of Dr. Sharda Sawant and Ms. Siddhi
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Redkar, Advanced Centre for Treatment, Research and Education in Cancer (ACTREC), Mumbai for TEM analysis and Microscopy Division, Carl Zeiss, India (Bangalore), Zeiss Group,
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PT
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M
AN
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for SEM analysis of the yeast samples.
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Table 1. Details of microorganis ms displaying high tolerance towards uranium Microorganism Tolerance/Minimum Site of isolation inhibitory concentration (MIC) (mM) Candida sorbophila 13 Uranium mine, Salamanca, Spain Acidithiobacillus 9 Uranium waste piles, ferrooxidans Johanngeorgenstadt, Germany Pseudomonas aeruginosa 5 Uranium ore deposit, DPs-13 Domiasiat, India Thermoterrabacterium 5 Hot springs, Yellow ferrireducens stone National Park, WY, U.S. Desulfovibrio desulfuricans 4 Wealden clay, G20 and I2 England Microbacterium oxydans 4 Radioactive waste depository, Russia Serratia marcescens 4 U ore deposit, Domiasiat, India Shewanella putrefaciens 200 3 Crude oil pipeline, Canada 3
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Rhodotorula mucilaginosa
Uranium mine, Salamanca, Spain
Reference
de Silóniz et al. (2002) Merroun and SelenskaPobell (2001) Sarma et al. (2014) Khijniak et al. (2005) Payne et al. (2002) Nedelkova et al. (2007) Kumar et al. (2011) Wade and DiChristina (2000) de Silóniz et al. (2002)
ACCEPTED MANUSCRIPT Table 2. Uranium biosorption capacities of different Microorga Bio Uraniu p Temper Equilib nism mass m H ature rium quan concent (ºC) time tity ration (min (mg) (mM) where not specifi ed) Halobacte 0.04 6. rium 0 noricense DSM15987 Geobacill 250 4.20 5. us 0 thermoleo vorans subsp strombolie nsis Bacillus 150 0.0042- 4. mojavensi 0.021 0s 7. 0
microorganisms Isotherm, Chemical Kinetic groups model, and/or Thermod molecules ynamic for U parameter binding s Phosphoryl, carboxyl
9.3
Bader et al. (2017 )
-
11
Özde mir and Kilinc (2012 )
-
-
48.2
-
-
-
50
Özde mir et al. (2017 a) Özde mir et al. (2017 b) Tsuru ta (2004 ) Li et al. (2016 )
4. 05. 0
-
-
60
-
-
9.04
3. 5
100
0.042
3. 0
Room temper ature
12 h
Freundlic h, pseudo second order
Amino, phosphate, carboxyl
>3.0
-
0.0289
7. 3
25
24 h
Langmuir , pseudo second order
-
84
0.1
AC
Streptomy 15 ces levoris
Streptomy ces sporoverr ucosus dwc-3 Arthrobac ter G975
0.00420.021
PT
100
CE
Bacillus vallismort is
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M
AN
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-
CR
IP
T
-
Load Refer ing ence (mg g-1 )
Carva jal et al. (2012 )
ACCEPTED MANUSCRIPT
Amycolato 0.00 psis sp. 1 K47
0.168
4. 0
40
Equilib rium time (min where not specifi ed) 150
Brachyba cterium sp. G1
0.04
6. 0
-
-
Myxococc 20 us xanthus
1
4. 5
28
10
Citrobacte r freudii
0.0840.63
-
Chemical groups and/or molecules for U binding
Load Refer ing ence (mg g-1 )
Langmuir ΔH° = 22.50 kJ mol-1 ΔS°= 0.014 kJ mol-1 K -
Carboxyl, hydroxyl, amide
38.8
US
-
Isotherm, Kinetic model, Thermod ynamic parameter s
T
Temper ature (ºC)
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p H
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Uraniu m concent ration (mM)
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Bio mass quan tity (mg)
-
Carboxyl
-
ED
M
Microorga nism
20-50
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CE
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25-55
Pseudomo nas aeruginos
3050
0.4
2. 4
22
-
Langmuir Carboxyl and Freundlic h, pseudo second order, ΔG0 = 6.54 kJ mol-1 spontaneo us, ΔH0 = 12.69kJ mol-1 endother mic, ΔS0 = 64.53 J mol-1 K -
Celik et al. (2018 )
971 Bader ± 29 et al. (2018 ) 28.5 Gonz 6 alezMuno z et al. (1997 ) 48 Xie et al. (2008 )
100
Hu et al. (1996
ACCEPTED MANUSCRIPT p H
Temper ature (ºC)
Equilib rium time (min where not specifi ed)
Chemical groups and/or molecules for U binding
Load Refer ing ence (mg g-1 )
) 7. 8
-
60
0.1
7. 8
-
24 h
Microcysti s aeruginos a
0.168
7. 0
-
Saccharo myces cerevisiae (heat killed) Rhodotoru 4 la glutinis (raw)
0.0042 30 5. 5
Langmuir
Amide, carboxyl
53.5
Achar ya et al. (2009 ) Achar ya et al. (2012 ) Li et al. (2004 )
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0.1
AN
ED
M
60
15
-
Exopolysaccharide, Polyphosphate bodies
77.3 5
Langmuir , Freundlic h isotherm -
-
44
Hydroxyl, carboxyl, phosphate, amino
29.8
Lu et al. (2013 )
Langmuir and Freundlic h Langmuir , pseudo second order, ΔG0 = 21.9 ± 0.1 kJ mol-1 spontaneo us, ΔH0 = 36.1 ±1.9
Amino, carboxyl
98.4
-
26
Bai et al. (2010 ) Bai et al. (2012 )
-
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0.588 25 6. 0
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Rhodotoru 5.15 la glutinis (magnetic ally modified)
Isotherm, Kinetic model, Thermod ynamic parameter s
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a strain CSU Synechoco ccus elongatus strain BDU/750 42 Anabaena torulosa
Uraniu m concent ration (mM)
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Bio mass quan tity (mg)
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Microorga nism
-
0.420 20-30 6. 0
-
30
ACCEPTED MANUSCRIPT Bio mass quan tity (mg)
Uraniu m concent ration (mM)
p H
Temper ature (ºC)
Equilib rium time (min where not specifi ed)
Isotherm, Kinetic model, Thermod ynamic parameter s
Chemical groups and/or molecules for U binding
kJ mol-1 endother mic, ΔS0 = 198.0± 6.3 J mol1 K irreversib le Sips Methyl, model, methylene pseudo first order, ΔG0 = 16.8 ± 0.1 kJ mol-1 spontaneo us, ΔH0 = 47.2± 3.7 kJ mol-1 endother mic, ΔS0 = 218 ± 12 J mol-1 K irreversib le Pseudo second order
Load Refer ing ence (mg g-1 )
-
30
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0.153 25-6. 0
4.20
28 4. 5
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Trichoder 0.42 ma 0 harzianum
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Rhodotoru 5 la glutinis (immobili zed)
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Microorga nism
-
Fusarium sp. #ZZF51
100
0.21
4. 0
Room Temp.
60
Lentinus sajor-caju MAFF 430306
-
0.840
4. 5
5-35
30
Langmuir , pseudo second order Freundlic h
17.3
Bai et al. (2014 )
196
Akhta r et al. (2007 ) Yang et al. (2012 ) Bayra moğlu et al.
Hydroxyl, carboxyl
16
Carboxyl, amino
128
ACCEPTED MANUSCRIPT Microorga nism
Bio mass quan tity (mg)
Uraniu m concent ration (mM)
p H
Temper ature (ºC)
Equilib rium time (min where not specifi ed)
Isotherm, Kinetic model, Thermod ynamic parameter s
Chemical groups and/or molecules for U binding
-
-
-
Padina sp. -
0.2104.20
4. 0
37
74
Carboxyl, sulfonate, hydroxyl
377
Cystoseria indica (calcium pretreated ) Cystoseria indica
1.47
4. 0
30
Langmuir , second order regressio n model Langmuir , pseudo second order
-
455
Khani et al. (2006 )
0.2104.20
4. 0
Langmuir , first order, Activatio n energy= -6.15 kJ mole -1 , ΔG0 = 8.7 kJ mole-1 spontaneo us, ΔH0 = -16.39 kJ mole-1 exothermi c, ΔS0 = 0.026 J mol-1 K reversible
-
198
Khani et al. (2008 )
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US
AN 3h
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M CE
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5. 06. 0
15
Phosphate
(2006 )
0.00210.5
PT
300330
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Schizophy llum commune
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(Alkali treated)
Load Refer ing ence (mg g-1 )
120150 h
280
Gunth er et al. (2014 ) Khani (2011 )
ACCEPTED MANUSCRIPT p H
Temper ature (ºC)
Catenella repens
50
0.420
4. 5
15-45
Equilib rium time (min where not specifi ed) 45
Isotherm, Kinetic model, Thermod ynamic parameter s
Chemical groups and/or molecules for U binding
Load Refer ing ence (mg g-1 )
Langmuir , Freundlic h, pseudo second order
Carboxyl, hydroxyl, sulphate, phosphate
303
T
Uraniu m concent ration (mM)
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Bio mass quan tity (mg)
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CE
PT
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AN
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Microorga nism
Bhat et al. (2008 )
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Table 3. Summary of the proteomic responses observed in different microorganis ms after exposure to uranium Microorganis Uranium Exposu Protein levels upProtein levels Referen m Concentrati re time regulated down-regulated ce on (h) (µM) Escherichia 50 2 NADH/quinone Adenine Khemiri coli MG1655 oxidoreductase phosphoribosyl et al. (WrbA) transferase (Apt), (2014) polynucleotide Escherichia 80 2 phosphorylase coli MG1655 Fe-Superoxide (PNPase), chaperon Khemiri dismutase (SodB), ClpB, HtpG et al. monooxygenases, (Hsp90) (2014) WrbA Catalase (KatE), Apt Caulobacter 500 3 Two component cell Aerobic oxidative Yung et crescentus signaling, amino acid phosphorylation, al. CB15N metabolism, phytase, cell motility, (2014) ABC transporter, chemotaxis, cell nucleoid associated adhesion, Electron protein, Structural transport maintenance of chain(ETC), chromosomes protein folding, S(SMC) proteins, cell layer biosynthesis growth and division associated proteins Anabaena 75 3 Phycobilisome core Phycobilisome core Panda PCC 7120 component (ApcF), membrane linker et al. phycobilisome rod (ApcE-1 and (2017) core linker protein ApcE-2), Enolase (CpcG1, (Eno), Fructose 1,6 CpcG4),phycocyanobil bisphosphate in lyase CpcS2 aldolase (Fda-1), (CpcS2-1), ribuloseFe-superoxide 1,5-bisphosphate dismutase(FeSOD), carboxylase/oxygenase PNPase, molecular , cysteine synthase A chaperones GroEL (CysA-1), (chaperonin), peroxiredoxins (Prx), DnaK (heat shock + ferredoxin-NADP protein) reductase
ACCEPTED MANUSCRIPT Exposu re time (h)
Protein levels upregulated
Protein levels down-regulated
Referen ce
3
Pyruvate dehydrogenase E1 subunit (Pde1), peroxiredoxins (Prx), ferredoxinNADP+reductase(F NR)
Panda et al. (2017)
Geobacter species
-
-
Geobacter sulfurreducen s DL-1
100
4
ATP synthase FOF1 subunit alpha (AtpA), beta (AtpB-1, AtpB-2), DNA directed RNA polymerase (RpoB-1), Mn-superoxide dismutase (MnSOD), transketolase (Tkt-b), DNA repair protein (UvrB) Ribosomal proteins, ATP synthase F1 subunit, tricarboxylic acid (TCA) cycle proteins (citrate synthase and isocitrate dehydrogenase), acetate-activating enzyme acetylcoenzyme A (CoA) hydrolase RND (resistancenodulation-cell division) family, lipid bilayer of outer membrane for protection, secretion systems, chaperones, c-type cytochromes, superoxide dismutase and reductase
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-
Wilkins et al. (2009)
Central metabolism, amino acid biosynthesis (GSU1061, GSU3099, GSU3095 and GSU1828), translation (GSU 1920), ribosome biogenesis (ObgE and EngB) Outer membrane protein 40 (Omp40)
Orellan a et al. (2014)
AC
CE
PT
ED
M
AN
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Anabaena L 31
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Uranium Concentrati on (µM) 200
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Microorganis m
Acidithiobacil lus ferrooxidans ATCC 23270
Microbacteri um oleivorans A9
500
-
10
24
Heat shock proteins(Hsps), cochaperone protein GrpE
Proteins involved in phosphate and iron metabolism, siderophores, DNA-
-
Dekker et al. (2016)
Gallois et al. (2018)
ACCEPTED MANUSCRIPT Exposu re time (h)
Protein levels upregulated
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AN
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binding transcriptional (ArsR family), DNA binding response regulator (NarL/FixJ family), Transporters (lipoprotein, Fe3+hydroxamate, dipeptide/oligopeptide/ nickel, enterochelin, Fe3+/spermidine/putres cine)
Protein levels down-regulated
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Uranium Concentrati on (µM)
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Microorganis m
Referen ce
ACCEPTED MANUSCRIPT Figure legends Fig. 1: Summary of the interactions of microorganisms with uranium including ‘omics’ based approaches.
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Fig. 2: Electron microscopy analysis of U challenged Yarrowia lipolytica cells. SEM images of
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(A) U untreated control cells or (B) cells treated with 50 µM uranyl carbonate for 24 h at pH 7.5. U untreated cells show characteristic oval morphology and bud scars (white arrows) (A) whereas
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the cells treated with uranium exhibited highly irregular and crustaceous surface morphology (B). TEM images of (C) U untreated cells or (D) 24 h U treated cells. Intracellular (black arrows)
M
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and extracellular (white arrows) uranyl deposits were observed in U treated cells.
Fig. 3: Illustration of the metabolic events exhibited by a filamentous cyanobacterium,
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Anabaena torulosa following exposure to 100 µM uranyl carbonate under phosphate limited
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conditions at pH 7.8 for 384 h (16 days). Bright field images were acquired by using Carl Zeiss Axioscop 40 microscope with oil immersion objectives (magnification x1,500, bars indicate 5
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µm). Red arrows indicate uranium association with polyphosphate bodies (24 h) and akinetes
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formed as a result of uranium stress (120 h).
Figure 1
Figure 2
Figure 3
Nilesh Kolhe, Smita Zinjarde, Celin Acharya PII: DOI: Reference:
S0734-9750(18)30126-5 doi:10.1016/j.biotechadv.2018.07.002 JBA 7277
To appear in:
Biotechnology Advances
Received date: Revised date: Accepted date:
15 February 2018 8 July 2018 9 July 2018
Please cite this article as: Nilesh Kolhe, Smita Zinjarde, Celin Acharya , Responses exhibited by various microbial groups relevant to uranium exposure. Jba (2018), doi:10.1016/j.biotechadv.2018.07.002
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ACCEPTED MANUSCRIPT Responses exhibited by various microbial groups relevant to uranium exposure Nilesh Kolhea,b,Smita Zinjardea*,c,Celin Acharyab*,d a
Institute of Bioinformatics and Biotechnology, Savitribai Phule Pune University, Pune 411007,
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India. b
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Molecular Biology Division, Bhabha Atomic Research Centre, Trombay, Mumbai 400085,
c
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India.
Department of Microbiology, Savitribai Phule Pune University, Pune 411007, India.
d
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Homi Bhabha National Institute, Anushakti Nagar, Trombay, Mumbai 400094, India.
*
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Authors for correspondence
S. Zinjarde address: Institute
of Bioinformatics
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Mailing
and
Biotechnology,
Savitribai Phule
Pune
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University, Pune 411007, India. Phone: + (91) 20 25691333, E-mail: [email protected];
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[email protected], Fax: + (91) 20 25690087 C. Acharya
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Mailing address: Molecular Biology Division, Bhabha Atomic Research Centre, Trombay,
25505326
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Mumbai 400 085, India. Phone: + (91) 22 25592256, E-mail: [email protected], Fax: + (91) 22
ACCEPTED MANUSCRIPT Abstract There is a strong interest in knowing how various microbial systems respond to the presence of uranium (U), largely in the context of bioremediation. There is no known biological role for uranium so far. Uranium is naturally present in rocks and minerals. The insoluble nature of the
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U(IV) minerals keeps uranium firmly bound in the earth’s crust minimizing its bioavailability.
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However, anthropogenic nuclear reaction processes over the last few decades have resulted in
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introduction of uranium into the environment in soluble and toxic forms. Microbes adsorb,
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accumulate, reduce, oxidize, possibly respire, mineralize and precipitate uranium. This review focuses on the microbial responses to uranium exposure which allows the alteration of the forms
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and concentrations of uranium within the cell and in the local environment. Detailed information on the three major bioprocesses namely, biosorption, bioprecipitation and bioreduction exhibited
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by the microbes belonging to various groups and subgroups of bacteria, fungi and algae is
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provided in this review elucidating their intrinsic and engineered abilities for uranium removal.
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The survey also highlights the instances of the field trials undertaken for in situ uranium bioremediation. Advances in genomics and proteomics approaches providing the information on
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the regulatory and physiologically important determinants in the microbes in response to
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uranium challenge have been catalogued here. Recent developments in metagenomics and metaproteomics indicating the ecologically relevant traits required for the adaptation and survival of environmental microbes residing in uranium contaminated sites are also included. A comprehensive understanding of the microbial responses to
uranium can facilitate the
development of in situ U bioremediation strategies. Keywords: Uranium, Microorganisms, Biosorption, Bioprecipitation, Bioreduction, Proteomics, Metaproteomics, Genomics, Metagenomics
ACCEPTED MANUSCRIPT 1. Introduction Anthropogenic activities related to nuclear processes such as mining, fuel processing, weapon production or nuclear accidents have resulted in contamination of the environment with uranium (U) (Garbisu and Alkorta, 2003; Markich, 2002). Uranium, with a density of 19 g cm-3 ,
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is a naturally occurring long-lived radionuclide that prevails in soils, rocks, seas and oceans
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(Gavrilescu et al., 2009). The most common isotopes of U in natural deposits are U-238, U-235
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and U-234 and they occur in the proportions of 99.27%, 0.72% and 0.0056%, respectively. Out
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of the three isotopes, U-238 is considered to be stable and has a half-life of 4.5 billion years (Krane, 1987). The average concentration of uranium in earth’s crust (naturally occurring) is
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reported to be 3 mg kg-1 (Kalin et al., 2005). In river water, uranium concentration ranges from 0.01-6.6 µg l-1 (depending on the strata) while 30 and 3.32 µg l-1 U have been reported in
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groundwater and seawater respectively under natural conditions (Markich, 2002). Very high
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concentrations of uranium have also been reported in some parts of the world as a result of
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various nuclear activities. For example, concentrations as high as 11.7 g l-1 in groundwater and 16 mg g-1 in soil and sediments have been reported at various United States Department of
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Energy (DOE) sites (Riley and Zachara, 1992). Nuclear waste storage sites at Oak Ridge
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National Laboratory, TN, U.S. are also reported to have high (0.8 mg g-1 ) contents of U (Kelly et al., 2007). The chemical guideline value for uranium in drinking water has been increased from 2 μg l-1 in 1998 to 15 μg l-1 in 2004 and subsequently to 30 μg l-1 in 2011 based on human studies from Sweden, Finland and USA (Ansoborlo et al., 2015). Uranium has two major oxidation states, +4 U(IV) and +6 U(VI) (Markich, 2002). The speciation of U depends on factors such as pH, redox potential, temperature, ionic strength, presence of other metals, suspended particles and dissolved CO 2 (Bernhard et al., 1998; Markich,
ACCEPTED MANUSCRIPT 2002; Lu et al., 2013). U(IV) is prevalent in ores in the form of uraninite [UO 2 (s)], that is insoluble and stable under anaerobic conditions. U(VI) on the other hand, prevails in aqueous systems under oxidizing conditions and speciation of this form is dependent on pH. Under acidic conditions (pH < 5), U(VI) exists in the soluble form as UO 2 2+ and is considered to be more toxic
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than U(IV) ions (Suzuki and Banfield, 2004). Hydroxides of uranium that predominate between
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pH 5.0 to 7.0 are less toxic towards aquatic flora and fauna (Markich, 2002; Choppin, 2007).
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Uranyl carbonate species such as UO 2 (CO 3 )2 2- and UO 2 (CO 3 )3 4- are predominant in rivers, ponds, lakes and sea water above circumneutral pH (>7.5) (Konstantinou and Pashalidis, 2004). 235
U) and chemical toxicity. Uranium
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Uranium, exhibits radiological (particularly from
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containing waste released into the environment can accumulate in the soil or move into aquatic systems like surface and groundwater (Gavrilescu et al., 2009). Consequently, uranium is likely
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to be introduced into terrestrial and aquatic food chains (Priest, 2001; Anke et al., 2009). Unlike
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some other metals, uranium does not have any biological role in living organisms. This metal has increased affinity for phosphate groups and binds strongly to phosphorylated peptides (Pardoux
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et al., 2012). In vivo studies have demonstrated that U treatment affects radical scavenging
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activity negatively in lung epithelial cells due to loss of glutathione (an effective antioxidant) and superoxide dismutase (SOD) enzyme (Periyakaruppan et al., 2007). U is known to cause hepatic,
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lung and renal damage in humans (Priest, 2001; Anke et al., 2009). Although many microbes have been suggested to respire uranium (Wade and Christina, 2000; Wagner et al., 2012), the toxicity of uranium in microbes has been established in terms of inhibition of microbial activity (Carvajal et al., 2012), distortion of cell surfaces and loss of cell viability (Sepulveda-Median et al. 2015), abortion of transcriptional and translational processes (Mukherjee et al., 2012), growth arrest and suspension of DNA replication (Park and Jiao, 2014) and oxidative damage (Volesky
ACCEPTED MANUSCRIPT and Holan, 1995). In order to resist uranium toxicity, the microbes employ various mechanisms which have shown potential for uranium immobilization and bioremediation. Uranium contamination poses substantial threat to human, environmental health and safety, therefore demanding a need to address the problem effectively. Physiochemical process
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used for treatment of metal contamination such as membrane filtration, chemical precipitation,
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electrochemical treatment, use of ion-exchange resins, incineration, excavation and extraction
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with organic solvents are often cost-intensive. Moreover, they are ineffective in removing toxic metals from large volumes leading to generation of hazardous by-products. The problems
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associated with physiochemical processes can be overcome by using microbial systems (Wang
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and Chen, 2006). Microorganisms interact with uranium and bring about alterations in the concentrations and/or oxidation states within the cell and their immediate environment thus
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regulating the mobility of uranium in the environment (Macaskie et al., 1992; Merroun and
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Seleneska-Pobell, 2008; Acharya, 2015; Majumder and Wall, 2017). Despite of the toxicity imposed by uranium, native microorganisms have shown tolerance and survival in uranium
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contaminated sites. Such isolates obtained from different parts of the world demonstrating
microorganisms
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significant uranium tolerance are summarized in Table 1. Various strategies adopted by for
persistent
existence
in
uranium
contaminated
environments
include
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mechanisms such as biosorption, biomineralization and biotransformation which can be used for in situ or ex situ uranium bioremediation (Fig. 1). Previously, reviews on uranium have discussed about speciation and bioavailability (Markich, 2002), bioreduction (Wall and Krumholz, 2006; Majumder and Wall, 2017), bacterial and cyanobacterial interactions (Pollmann et al., 2006; Merroun and Seleneska-Pobell, 2008; Acharya and Apte, 2013b) and biogeochemistry and bioremediation of soil ( Gavrilescu et al.,
ACCEPTED MANUSCRIPT 2009; Newsome et al., 2014; Selvakumar et al., 2018). However, to the best of our knowledge, there is no comprehensive recent review detailing uranium interactions with different microbial groups and subgroups of prokaryotes (archaea and bacteria-Gram positive/Gram negative) and
Basidiomycetous Rhodophyta).
and
We
and
filamentous
Zygomycetous
discuss
in
groups
detail three
fungi and
sub-categorized
into
algae-Chlorophyta,
microbial processes
Ascomycetous, Phaeophyta
and
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(fungi-yeasts
important for uranium
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eukaryotes
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bioremediation namely, biosorption, bioprecipitation and bioreduction with respect to various microbial groups (Fig. 1). Subsequently, the role of natural and genetically engineered strains in
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mediating such interactions is discussed. We conclude by detailing the use of ‘omics’ based
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approaches for understanding uranyl interactions at the molecular level. 2. Biosorption of uranium
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Biosorption is a property of live or dead biomass to adsorb and concentrate toxic metals
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from very dilute aqueous solutions including uranium (White and Gadd, 1990; Volesky and Holan, 1995). In the recent years, biosorption has emerged as an important alternative for
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removing radionuclides from aqueous solutions (Lloyd and Macaskie, 2002). This phenomenon is
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often rapid and growth-independent since metabolic processes are not involved. Biomass generated from different biotechnological and agricultural processes can be used as biosorbents
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presenting the economic viability of the biosorption process (Ahluwalia and Goyal, 2007). This method often displays high efficiency and selectivity for metals over a broad range of pH and temperature. A summary of important parameters studied during the process of microbial biosorption of uranium are included in Table 2. Different microbial groups that have been used for biosorption of uranium are detailed in the following sections. Few examples of bioaccumulation have also been included in this section. Bioaccumulation is sometimes referred interchangeably
ACCEPTED MANUSCRIPT with biosorption wherein the accumulated metal is located intracellularly in the metabolically active cells. 2.1. Biosorption of uranium by archaea and bacteria Archaea and bacteria are ubiquitous and often prevail in extreme environments (Battista,
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1997; Selenska-Pobell et al., 2001; Jroundi et al., 2007; Acharya et al., 2012). They have high
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surface to volume ratios and different functional groups (phosphate, carboxyl, amide and
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hydroxyl) present on their surfaces which complex with metals and radionuclides (Vieira and Volesky, 2000; Bader et al. 2018). Archaea and bacteria differ in cell wall characteristics. Unlike
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bacteria, archaea lack peptidoglycan in their cell walls.
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2.1.1. Biosorption of uranium by archaea
The role of archaea belonging to genus Halobacterium in bringing about uranium
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biosorption has been recently demonstrated. Two strains of Halobacterium noricense isolated
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from Waste Isolation Pilot Plant repository were able to associate with uranium in a multistage process. An initial phase of rapid sorption was followed by uranium release and subsequently, a
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slow re-association of uranium with the cells was observed. Carboxylic and phosphoryl groups
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were important in uranyl binding during the first phase of biosorption (Bader et al., 2017; 2018). 2.1.2. Biosorption of uranium by Gram positive bacteria
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Members of the Bacillus, Geobacillus, Clostridium, Streptomyces and Arthrobacter have been shown to function as efficient uranium biosorbents. Bacillus species have been used for biosorption of uranium on several occasions. The site specific complex model (SCM) initially described by Fein et al. (1997) has been used for quantitative prediction of uranyl adsorption onto the cells of Bacillus subtilis under different conditions (Fowle et al., 2000; Gorman-Lewis et al., 2005). Recently, the efficacy of alginate-chitosan microcapsules with immobilized B.
ACCEPTED MANUSCRIPT subtilis in uranium sorption has been demonstrated (Tong, 2017). Proteinaceous surface layer (Slayer) representing the outermost cell envelope component of bacteria has been implicated in uranium biosorption. In Bacillus sphaericus JG-A12 (isolated from uranium mining waste pit), the S-layer was shown to bind uranium via phosphate and carboxyl groups (Raff et al., 2002;
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Merroun et al., 2005). Cells, spores and S-layer of this strain embedded in silica gel displayed
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promising potential for uranium adsorption from wastewater samples (Raff et al., 2002, 2003).
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Extended X-ray absorption fine structure (EXAFS) analysis suggested the coordination of U with phosphate groups in meta-autunite mineral phase (Merroun et al., 2005). Pelosinus sp. strain
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UFO1 [native to Oak Ridge Field Research Centre (FRC), TN, U.S.] showed potential for
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sequestering uranium in two different oxidation states, U(IV) and U(VI) (Ray et al., 2011). Furthermore, the uranium-binding complex (UBC) of strain UFO1 comprising of two S-layer
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domain proteins was shown to bind U(IV) (Thorgersen et al., 2017) in contrast to S-layer of
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Bacillus sphaericus JG-A12 which bound U(VI) (Merroun et al. 2005). Detailed analysis of uranium biosorption in Bacillus sp. (dwc-2) obtained from a radioactive waste disposal site in
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Southwest China revealed the involvement of ion-exchange, complexation and bioaccumulation
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process (Li et al., 2014). Biomass based solid-phase extraction of metals from environmental samples has emerged as a promising technique owing to its simplicity and low application cost
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(Özdemir et al., 2013). Cells of Bacillus mojavensis and Bacillus vallismortis loaded onto multiwalled carbon nanotubes were employed as solid-phase extractors to pre-concentrate uranium from aqueous solutions and lake water samples (Özdemir et al., 2017a, 2017b). In batch process at pH 5.0, multiwalled carbon nanotubes with immobilized B. vallismortis could load up to 25.8 mg U g-1 and could be used for 30 cycles of adsorption and desorption without significant decrease in biosorption efficiency. In yet another study, a thermophilic bacterium, Geobacillus
ACCEPTED MANUSCRIPT thermoleovorans subsp. stromboliensis immobilized on Amberlite XAD-4 ion-exchange resin showed U sorption of 11 mg U g-1 at pH 5.0 (Table 2) (Özdemir and Kilinc, 2012). Ionic Liquids (ILs) are commonly used to concentrate and extract radionuclides from waste streams (Binnemans, 2007; Rout et al., 2012). Certain ILs (1-butyl-3-methylimidazolium
T
hexafluorophosphate [BMIM] [PF6 ], N-ethylpyridinium trifluoroacetate [EtPy] [CF 3 COO] and
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N-ethylpyridinium tetrafluoroborate [EtPy] [BF 4 ]) were shown to decrease U biosorption by
CR
Clostridium sp. due to membrane damage of the cells stressing the need to formulate biocompatible ionic liquids for efficient microbial U recovery (Zhang et al., 2014).
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Several actinomycetes have shown the promising potential for uranium adsorption (Friis
AN
and Myers-Keith, 1986; Tsuruta, 2004). For example, Streptomyces levoris demonstrated essential properties for biosorption of uranium from acidic solutions which was mediated via
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phosphate groups of teichoic acids harbored in the cell walls. These groups conferred a net
ED
negative charge on bacterial cells and allowed strong binding of positively charged uranyl ions in pH range of 3.5-6.0 (Tsuruta, 2004). Similarly, the phosphodiester residues of cell wall and
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cytoplasmic fraction of Streptomyces longwoodensis bound uranium with appreciable efficiency
CE
(Friis and Myers-Keith, 1986). Recent studies on Streptomyces sporoverrucosus dwc-3 isolated from a radioactive waste disposal site in China have revealed the binding of U to amino,
AC
phosphate and carboxyl groups of the cell walls at pH 3.0 (Li et al., 2016). Various Arthrobacter species included within actinomycetes have been reported for uranium biosorption. Arthrobacter strains are abundantly found in uranium rich environments (Updegraff and Douros, 1972; Miller et al., 1987; Suzuki and Banfield, 2004; Chapon et al., 2012; Kumar et al., 2013a). Arthrobacter ilicis accumulated uranium intracellularly as precipitates within polyphosphate granules (Suzuki and Banfield, 2004). Reduced uranium adsorption was observed in Arthrobacter species G975 in
ACCEPTED MANUSCRIPT the presence of aqueous bicarbonate, a competing ligand for U(VI) (Carvajal et al., 2012). Uranium mobility is governed in aquatic systems by the formation of highly soluble and stable uranyl carbonate complexes namely, UO 2 CO 3 0 , UO 2 (CO 3 )2 2- and UO 2 (CO 3 )3 4- above circumneutral conditions (Langmuir, 1978; Guillaumont et al., 2003). These uranyl species
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interfere with uranium complexation on microbial surfaces. Two more actinomycetes namely
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Amycolatopsis sp. K47 and Brachybacterium sp. G1 have recently been shown to remove
related to these groups of bacteria are detailed in Table 2.
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2.1.3. Biosorption of uranium by Gram negative bacteria
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uranium via biosorption (Celik et al., 2018; Bader et al., 2018). Important biosorption parameters
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Lipopolysaccharide (LPS) and cell wall components play an important role in metal binding in Gram negative bacteria. Phospholipids and phosphate groups of LPS are particularly
M
significant for binding of metal ions (Beveridge and Fype, 1985). Different types of Gram
ED
negative bacteria exhibiting U adsorption are catalogued here. Two members of the family Enterobacteriaceae namely, Citrobacter and Serratia have
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shown prominence for their uranyl sorption abilities (Xie et al., 2008; Kumar et al., 2011).
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Adsorption of U(VI) by Citrobacter freudii was found to be rapid with carboxyl groups playing an important role in uranium binding (Xie et al., 2008). Dead biomass of this bacterium
AC
displayed better efficiency than live cells indicating a metabolism-independent process of uranium sorption. A strain of Serratia marcescens isolated from sub-surface soil of a uranium ore deposit exhibited high tolerance towards uranium (4 mM) under acidic conditions (pH 3.5). This bacterium removed 92% (21-22 mg l-1 ) and 60-70% (285-335 mg l-1 ) of uranium from 100 µM and 2 mM uranyl solutions, respectively (Kumar et al., 2011). Similarly, an extremely uranium tolerant strain, Cupriavidus metallidurans CH34, resisting up to 30 mM uranium,
ACCEPTED MANUSCRIPT immobilized U(VI) by complexing it with phosphoryl or carboxyl containing biomolecules of LPS layer (Llorens et al., 2012). Uranium adsorption onto bacterial cell surface is largely dependent on pH. Uranium speciation and characteristics of bacterial cell wall change as the function of pH. The uranyl
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content adsorbed onto Shewanella oneidensis MR-1 below pH 5.0 was unaffected by dissolved
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inorganic carbon (DIC) concentrations due to the absence of uranyl carbonate complexes.
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However, at higher pH, the extent of uranium adsorption decreased with increasing concentrations of DIC because of the formation of anionic uranyl-hydroxide-carbonates on bacterial surfaces
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(Sheng and Fein, 2013).
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Members of the genus Pseudomonas (belonging to the Gammaproteobacteria group) have also been studied with respect to uranium biosorption (Hu et al., 1996; Choi and Park, 2005; Choi
M
et al., 2009). The live or dead cells of Pseudomonas putida exhibited higher specificity for uranyl
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adsorption when exposed to mixed metal waste solutions comprising of uranium (U), lead (Pb), cadmium (Cd), zinc (Zn) and nickel (Ni) at pH 6.0 (Choi et al., 2009). Compared to soil and clay
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constituents, dead cells of P. putida were more effective in uranium adsorption (Choi and Park,
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2005). The cells of Pseudomonas aeruginosa CSU showed rapid intracellular accumulation of uranium within < 10 s of exposure (Strandberg et al., 1981). The same organism demonstrated
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uranyl adsorption in the presence of transition metals under acidic and neutral conditions equivalent to commercial cation-exchange resins. However, certain cations [ferric iron (Fe 3+) acting as uranium analogues] inhibited binding of U to P. aeruginosa biomass in a significant manner (Hu et al., 1996). Removal of iron from wastewater was therefore proposed as a prerequisite step for efficient utilization of P. aeruginosa CSU biomass for uranium adsorption.
ACCEPTED MANUSCRIPT Uranyl adsorption onto Myxococcus xanthus (belonging to Deltaproteobacteria) was demonstrated to be dependent on culture age, concentration of uranyl ion and pH. Older cultures (72 h) when exposed to 1 mM uranium showed maximal adsorption (2.4 mM of U g-1 of dry biomass) at pH 4.5 (Table 2). Adsorbed uranium was found to be distributed on the cell wall and
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extracellular mucopolysaccharide (Gonzalez-Munoz et al., 1997). Another strain of M. Xanthus
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demonstrated uranium coordination with cell surface-associated organic phosphate groups at very
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low pH (2.0) (Jroundi et al., 2007).
Cyanobacteria are morphologically diverse, oxygenic, photosynthetic, Gram negative
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bacteria commonly present in freshwater, marine and terrestrial habitats (Acharya and Apte
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2013b). They are capable of tolerating, accumulating and detoxifying metals in water bodies thereby altering their mobility and bioavailability (Parker et al., 2000; Pradhan and Rai, 2000;
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Zhou et al., 2004; Acharya et al., 2013). Majority of the uranyl binding studies with
ED
microorganisms have been performed in the pH range of 1.5 to 6.0 wherein, the uranyl ion (UO 2 2+) and hydroxo complexes like UO 2 (OH)+ and (UO 2 )2 (OH)2 2+ predominate (Markich,
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2002; Choppin, 2007). The inability of several microorganisms to sequester U above pH 7.0 is
CE
attributed to the repulsion between anionic carbonate complexes of U such as [UO 2 (CO 3 )2 2-] and [UO 2 (CO 3 )3 4-] and negatively charged cell surfaces (Markich, 2002; Konstantinous and
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Pashalidis, 2004; Acharya and Apte, 2013b). However, two marine cyanobacteria have demonstrated the ability to remove significant quantities of uranium from aqueous solutions above pH 7.0 (Acharya et al., 2009, 2012, 2013; Acharya and Apte 2013a). A unicellular marine cyanobacterium, Synechococcus elongatus showed rapid binding (53.5 mg g-1 dry weight of U) within 5-10 min when exposed to 100 µM U supplemented as uranyl carbonate at pH 7.8 (Acharya et al., 2009). Most of the bound uranium complexed with amide and deprotonated
ACCEPTED MANUSCRIPT carboxyl groups of extracellular polysaccharides (EPS) could be released by HCl or EDTA treatment. The organism tolerated 0.5 M NaCl and demonstrated high uranyl loading (2960 µg g1
) from simulated sea water in 4 weeks under continuous replenishment conditions (Acharya et
al., 2013). Immobilized cells of S. elongatus were also effective for U binding under continuous conditions
(Acharya
and
Apte,
2013b).
Regeneration
of
immobilized
T
flow-through
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Synechococcus biomass for multiple adsorption-desorption cycles (up to 3 times) was observed
uranium by S.
elongatus,
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above pH 7.0 without any significant loss of U binding (90-92%). In contrast to rapid binding of another marine cyanobacterium, namely Anabaena torulosa
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demonstrated rather slow accumulation of uranium in polyphosphate bodies when the cells were
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exposed to 100 µM uranyl carbonate at pH 7.8 for 24 h (Acharya et al., 2012, 2013a). A naturally abundant bloom-causing cyanobacterium, Microcystis aeruginosa, exhibited potential
M
for uranyl adsorption between pH 4.0 to 8.0. Interestingly, uranyl adsorption by this organism
ED
increased when pH was shifted from 10.0 to 11.0 due to a decrease in intrinsic viscosity of the slime suspension at higher pH (Li et al., 2004).
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2.2. Biosorption of uranium by yeasts and filamentous fungi
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Yeasts and filamentous fungi are lower eukaryotes that can be easily cultivated to produce large quantities of biomass. They are amenable to genetic manipulation and have several
AC
biotechnological and industrial applications (Volesky and Holan, 1995; Bai and Abraham, 2003; Zinjarde, 2014; Zinjarde et al., 2014). Fermentation processes involving yeasts and fungi result in the generation of large quantities of relatively inexpensive biomass. This biomass has been used for effective biosorption of toxic metals and radionuclides (Kapoor and Viraraghavan, 1995; Wang and Chen, 2006). The role of yeasts and filamentous fungi in biosorption of uranium from aqueous solutions has been discussed in the following sections.
ACCEPTED MANUSCRIPT 2.2.1. Biosorption of uranium by yeasts Easy cultivation on a large scale with high yields and generation of waste biomass from food and beverage industries are some of the advantages of yeasts which present them as promising biosorbents for heavy metal decontamination (Volesky and Holan, 1995). Uranyl
T
adsorption by the ascomycetous yeast Saccharomyces cerevisiae has been reported on several
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occasions (Strandberg et al., 1981; Liu et al., 2010; Faghihian and Peyvandi, 2012; Lu et al.,
CR
2013). Dead cells of S. cerevisiae demonstrated higher metal biosorption capacity compared to the live cells (Soares et al., 2002; Xia et al., 2013). The heat killed cells of S. cerevisiae
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apparently released phosphate due to membrane damage which complexed with uranyl residues
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and formed uranyl phosphate nanoparticles on the cell surface (Wang et al., 2017). Amidoxime groups, known for their selectivity for uranium were successfully grafted on S. cerevisiae
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biomass and used for uranyl adsorption from salt lake brine solutions with good efficiency (Bai
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et al., 2016). Dead biomass of S. cerevisiae with rough surfaces and ‘nano-holes’ provided larger contact area for uranyl sorption as compared to the live cells with smooth surfaces (Wang et al.,
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2017). A marine strain of the ascomycetous yeast, Yarrowia lipolytica isolated from oil-polluted
CE
seawater (Zinjarde and Pant, 2002) is currently being explored in the context of uranyl removal (Kolhe et al., unpublished data). Preliminary results have demonstrated the ability of this strain
AC
to sequester 50% of the input uranium (50 µM U) loading up to 37.51 mg U g-1 dry weight of cells at pH 7.5. The cellular localization of the uranyl deposits in the yeast cells was examined by electron microscopy. Scanning electron microscopy (SEM) analysis of the cells exposed to 50 µM uranyl carbonate for 24 h revealed irregular, crustaceous surface in contrast to control, U untreated cells which showed smooth cell surfaces (Figs. 2A and 2B). Transmission electron microscopy (TEM) of U exposed yeast cells revealed electron dense, crystalline needle/spicule
ACCEPTED MANUSCRIPT like deposits on cell surfaces and at intracellular locations (Fig. 2D) as compared to control, U untreated cells which did not show any such crystalline precipitates (Fig. 2C) establishing uranyl association with the yeast cells. Metals interact with different functional groups of proteins, lipids or carbohydrates
T
present on cell walls of the biomass (Chen et al., 2007). Chemical modification of these groups
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can modulate biosorption processes. For example, esterification of carboxyl groups and
CR
methylation of the amino groups in the cell wall of the basidiomycetous yeast, Rhodotorula glutinis by methanol and formaldehyde treatment respectively, enhanced uranyl adsorption as
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compared to the chemically unmodified cells (Bai et al., 2009; 2010). Biomass is usually
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recovered by centrifugation or filtration after adsorption which adds to the treatment costs. ‘Smart biosorbents’ resulting from introduction of magnetic phase into the microbes allowed
M
metal removal and biomass recovery by applying external magnetic fields (Wong and Fung,
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1997; Wang et al., 2000; Yavuz et al., 2006; Rao et al., 2013). Magnetically modified biomass of R. glutinis has been successfully utilized for uranium biosorption (Bai et al., 2012).
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Immobilization of microbial biomass imparts mechanical strength, confers chemical stability and
CE
allows regeneration of the biomass. Calcium-alginate immobilized biomass of R. glutinis showed improved U removal from 0.74 to 17.3 mg g-1 U when the input uranyl concentration was
AC
increased from 2.5 to 40 mg l-1 , respectively (Bai et al., 2014). Various parameters related to biosorption of uranium by raw and modified biomass of R. glutinis are detailed in Table 2. Recently, another species of Rhodotorula namely, R. mucilaginosa BII-R8 is reported in context of uranium biosorption (U-hydroxides and U-hydroxo-carbonates) by means of carboxyl and phosphate groups of the biomass (Lopez-Fernandez et al., 2018). 2.2.2. Biosorption of uranium by filamentous fungi
ACCEPTED MANUSCRIPT Filamentous fungi have high area to mass ratios and their biomass has been used for the removal of metal ions from the aqueous solutions (Kapoor and Viraraghavan, 1995). Filamentous fungi belonging to the ascomycetous group have been studied for biosorption of uranium as discussed here. Viable and non-viable biomass of Trichoderma harzianum has
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demonstrated better uranyl adsorption capacities than commercial resins such as Dowex-SBR-P
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and IRA-400 (Akhtar et al., 2007). This biomass could be used over five cycles of adsorption-
CR
desorption thereby presenting its cost-effectiveness for uranyl removal. Calcium alginate immobilized biomass of T. harzianum showed better biosorption efficiency than free cells
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(Akhtar et al., 2009). Higher values of sorption (87-97%) were observed for immobilized cells as
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compared to free cells which demonstrated 40-75% when exposed to 100-400 mg l-1 of input U within 8 h (Akhtar et al., 2009). Similarly, another ascomycetous fungus, Aspergillus fumigatus
M
immobilized in calcium alginate, was used for biosorption of uranium with appreciable
ED
efficiency (Wang et al., 2010). Chemical modification by formaldehyde, methanol and acetic acid enhanced affinity of a mangrove endophytic fungus, Fusarium sp. #ZZF51 towards uranium
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(Chen et al., 2014). Non-viable cells of two air-borne, uranium and acid tolerant fungi,
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Aphanocladium spectabilis and Acremonium minutisporum, showed maximum uranium sorption capacities of 162.1 and 161.5 mg U g-1 of dried biomass, respectively (Gargarello et al., 2008).
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Basidiomycetous fungi have also been implicated for removal of uranium from aqueous solutions. Lentinus sajor-caju is a saprophytic basidiomycetous white-rot fungus that grows easily on carbon sources such as cellulose and produces several extracellular enzymes for bioremediation of xenobiotics compounds (Rodrıguez et al., 2004). The alkali treated biomass of this fungus was found to be more effective in removing uranium than its untreated form (Bayramoğlu et al., 2006). Improved sorption capacity of alkali-treated biomass was apparently
ACCEPTED MANUSCRIPT due to increased availability of binding sites resulting from deacetylation of chitin to chitosan. Another basidiomycetous fungus
Schizophyllum
commune
showed
uranium accumulation
intracellularly within the vacuoles and on the cell walls (Gunther et al., 2014). Unlike other uranyl-microbial interactive studies which demonstrated a predominant role of carboxyl groups
T
in uranium complexation (Merroun et al., 2005; Acharya et al., 2009), the phosphate groups
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rather than the carboxyl groups facilitated the uranyl adsorption in S. commune.
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Rhizopus arrhizus, a zygomycetous fungus has also been explored for biosorption of uranium from aqueous solutions (Tsezos and Volesky, 1982; Tsezos, 1983). In an attempt to
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evaluate the competing ion effect, it was observed that aluminiun did not interfere in the
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kinetics of uranyl adsorption in R. arrhizus (Tsezos et al., 1996). Certain products derived from fungi like chitin and chitosan have also shown the ability for biosorption of uranium
Algae
are
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2.3. Biosorption of uranium by algae
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(Sakaguchi et al., 1981; Wang et al., 2009).
photosynthetic,
autotrophic
organisms
abundantly
found
in
aquatic
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environments that can serve as inexpensive biosorbent material. Biosorption of uranium by
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different groups of algae belonging to chlorophyta (green), phaeophyta (brown) and rhodophyta (red) groups has been discussed in the following sections.
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2.3.1. Biosorption of uranium by green algae Green algae are a diverse group of photosynthetic microorganisms ranging from unicellular flagellates to complex multicellular forms. Green algae belonging to the genera Chlorella, Spirulina and Nostoc have been effective in the removal of uranium ions from aqueous solutions. Live and dead cells of Chlorella vulgaris brought about rapid biosorption of uranium between pH 3.0 to 6.0 in mineral medium with low phosphate concentration (Vogel et
ACCEPTED MANUSCRIPT al., 2010). However, after prolonged incubation (48 to 96 h), live cells caused the mobilization of bound uranium by releasing organic acids. C. vulgaris did not adsorb uranium from natural or artificial sea water at pH 8.0. The observed inhibition of uranyl uptake was due to the formation of stable uranyl-carbonate complexes, [UO 2 (CO 3 )2 2-] and [UO 2 (CO 3 )3 4-] (Nakajima et al., 1979).
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In contrast, uranyl adsorption by C. vulgaris immobilized in polyacrylamide gels was unaffected
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by pH. The immobilized cells could recover 100% U from sea water spiked with 10 mg l-1 U
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(Nakajima et al., 1979). In yet another investigation, uranium (UO 2 2+ ions) solubilized from
and Nostoc linckia (Cecal et al., 2012).
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2.3.2. Biosorption of uranium by brown algae
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uranium ores and sludges were found to be sequestered within the cells of Spirulina platensis
Brown algae are multicellular in nature and are prevalent in marine ecosystems. The
M
brown color displayed by this group of algae is due to the presence of fucoxanthin (Groisillier et
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al., 2014). The biomass of Sargassum fluitans could adsorb up to 105 mg g-1 of uranium in 30 days under continuous flow through conditions (Yang and Volesky, 1999). The biomass of
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Padina sp. and Padina pavonia has been demonstrated for uranyl sorption (Khani, 2011; Aytas
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et al., 2014). Cystoseria indica has also demonstrated good performance for uranyl biosorption under acidic conditions from pH 2.5 to 4.0 (Table 2) (Khani et al., 2006; Khani et al., 2008;
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Ghasemi et al., 2011; Gök et al., 2017). Chemically modified biomass of Cystoseira sp. did not affect the process of uranium adsorption from simulated wastewater samples containing different metal ions. The biosorption process was exothermic and spontaneous as indicated by negative values of enthalpy and free energy changes (Gök et al., 2017). Alginate, a polysaccharide, derived from some brown algae has also been used for biosorption of U from aqueous solutions (Gok and Aytas 2009; Yu et al., 2017).
ACCEPTED MANUSCRIPT 2.3.3. Biosorption of uranium by red algae Natural occurrence of red algae on sea coasts and estuarine waters in large quantities allows this biomass to be used as an inexpensive biosorbent. The red alga Catenella repens showed encouraging results for biosorption of uranium under acidic conditions displaying
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maximum loading capacity of 303 mg U g-1 from solution containing 100 mg l-1 U (Bhat et al.,
ores and sludges (Cecal et al., 2012). 3. Bioprecipitation of uranium offers
a
viable
remediation
technique
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Bioprecipitation
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2008). Porphyridium cruentum, another red alga, has been shown to remove UO 2 2+ ions from
for
controlling
metal
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contamination. Microbial bioprecipitation of metals is often an enzymatic process. Phosphatases produced by microorganisms hydrolyze organic phosphate substrates and release inorganic
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phosphate (Pi). The latter interacts with metals in solutions and precipitates them in the form of
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insoluble minerals, generally metal phosphates (Lunt et al., 2007). Therefore, bioprecipitation is often referred as biomineralization as the end product is an insoluble mineral.
Alternatively, the
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degradation of polyphosphate (phosphate polymer) resulting in phosphate release and metal
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bioprecipitation has also been reported in microbes (Renninger et al. 2004; Acharya et al. 2017). Several environmental bacteria isolated from uranium enriched or contaminated sites have been
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shown to precipitate uranium as uranyl phosphate owing to their constitutive phosphatase activity (Martinez et al., 2007; Beazley et al., 2007; Merroun et al., 2011; Newsome et al., 2015). The crystalline uranium precipitates resulting from bioprecipitation of U show no alteration in the U(VI) redox state in contrast to reductive precipitation by dissimilatory metal reducing bacteria under anaerobic conditions (Wall and Krumholz, 2006). U immobilization as U(VI) phosphate (autunite) minerals under aerobic conditions can be utilized for alleviating uranium at
ACCEPTED MANUSCRIPT contaminated sites for extended periods over a wide pH range as they are not susceptible to changes in oxidation conditions. The following section highlights the potential of different microbial groups for bioprecipitation of uranium (Fig. 1). 3.1. Bioprecipitation of uranium by bacteria
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Bacteria, owing to their diversity and distinct metabolic properties, are capable of
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modulating biogeochemistry of metals. A plethora of bacteria from different geographic locations
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can precipitate aqueous uranium in the form of insoluble uranium phosphates as detailed in the following part of the review.
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3.1.1. Bioprecipitation of uranium by Gram positive bacteria
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Gram positive bacteria belonging to the genus Bacillus isolated from radionuclide and metal polluted terrestrial locations have demonstrated uranium precipitation via constitutive acid
M
phosphatase activity (Beazley et al., 2007; Martinez et al., 2007). These studies involved the use
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glycerol-3-phosphate (G3P) as a substrate for enhancing phosphatase activity (Fig. 1). Bioprecipitation of U(VI) in Bacillus sphaericus JG-7B (isolated from uranium mining waste
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sediment) at pH 3.0 and 4.5 occurred via acid phosphatases in the absence of organic phosphate
CE
substrates. Phosphorylated biomolecules such as nucleic acids that were released from dead cells were proposed to serve as substrates for the phosphatases (Merroun et al., 2011). In most of the
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cases, uranium was precipitated in the form of stable and insoluble autunite/meta-autunite mineral phases (Merroun et al., 2006; Beazley et al., 2007; Martinez et al., 2007; Merroun et al., 2011; Theodorakopoulos et al., 2015). A facultative anaerobic bacterium Paenibacillus sp. JGTB8 isolated from uranium mining waste pile mineralized U under aerobic conditions (Reitz et al., 2014). Under anaerobic settings, mineral formation in JG-TB8 was not observed at pH 4.5 or 6.0 due to significant suppression of the constitutive phosphatase activity.
ACCEPTED MANUSCRIPT Microbacterium oxydans isolated from Siberian radioactive waste depository precipitated uranium extracellularly at pH 4.5 whereas uranium was bound to organophosphate groups such as fructose 6 phosphates at pH 2.0 (Nedelkova et al., 2007). A Microbacterium strain isolated from Chernobyl soil sample exhibited high tolerance to U and demonstrated intracellular
mechanism (Theodorakopoulos et al., 2015).
are
several
Alphaproteobacteria,
reports
on
various
Gammaproteobacteria
Gram
and
negative
bacteria
Deltaproteobacteria
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There
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3.1.2. Bioprecipitation of uranium by Gram negative bacteria
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accumulation in the form of autunite mineral formed as a result of phosphate released by active
belonging
classes
to
exhibiting
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bioprecipitation of uranium which are detailed in this section.
Two Alphaproteobacteria namely, Sphingomonas sp. S15-S1and Caulobacter crescentus
M
exhibited uranium precipitation by the virtue of constitutive acid and alkaline phosphatase
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activities, respectively (Merroun et al., 2006; Yung and Jiao, 2014). phoY gene encoding the periplasmic enzyme, PhoY (responsible for whole cell phosphatase activity) identified in the
PT
annotated genome of C. crescentus facilitated uranyl precipitation. A PhoY deletion mutant of C.
CE
crescentus neither released Pi in presence of glycerol-2-phosphate (G2P) nor precipitated U. The results established the role of PhoY in uranium precipitation (Yung and Jiao, 2014). (reclassified
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Citrobacter
later
as
Serratia,
Pattanapipitpaisal et
al.,
2002)
a
Gammaproteobacterium in the presence of glycerol phosphate was able to bioprecipitate uranium (Macaskie et al., 1992; Macaskie et al., 1994). Uranyl phosphate mineral phases are considered to be very stable and are not amenable to oxidative remobilization unlike the products of uranium reduction which can be re-oxidized to mobile U (Senko et al., 2002). Detailed analysis of uranium precipitation in Citrobacter revealed that monophosphate groups harbored
ACCEPTED MANUSCRIPT within the lipopolysaccharide (LPS) layer facilitated the initial nucleation. Further consolidation and growth of crystalline uranyl phosphate mineral was achieved by enzymatically generated phosphate groups (Macaskie et al., 2000). Expression of higher levels of phosphatases alone did not result in uranyl precipitation (Basnakova et al., 1998). Rather, a combination of cell surface
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architecture and phosphatase overproduction facilitated efficient U precipitation. Whole cells of
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Citrobacter immobilized in polyacrylamide matrices demonstrated impressive U loading of 9 g
CR
U g-1 of dry weight in the form of uranyl phosphate as a result of phosphatase mediated U precipitation (Macaskie et al., 1991). Most of the studies on uranium bioprecipitation have been
US
demonstrated under aerobic conditions. In contrast, a study involving Serratia sp. isolated from
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sediments of a nuclear site exhibited uranyl precipitation under anaerobic conditions (Newsome et al., 2015). Constitutive phosphatase activity and uranium bioprecipitation in yet another
M
Serratia strain were uncompromised even after exposure to high levels of ionizing radiation
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presenting its utility for nuclear waste decontamination (Paterson‐Beedle et al., 2012). Similarly, other members of Gammaproteobacteria such as Acidithiobacillus ferrooxidans, Rahnella sp.,
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Pseudomonas aeruginosa, Rhodonobacter, Acinetobacter sp. and Pseudomonas fluorescens and
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Deltaproteobacteria member like Myxococcus xanthus have demonstrated uranyl precipitation via Pi derived through phosphatase activity or polyphosphate hydrolysis (Fig. 1) (Merroun et al.,
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2003; Renninger et al., 2004; Jroundi et al., 2007; Beazley et al., 2007; Choudhary and Sar, 2011; Sousa et al., 2013; Sowmya et al., 2014; Krawczyk-Bärsch et al., 2015). In addition to the reports on uranium bioprecipitation by planktonic cells, biofilms of P. fluorescens precipitated U in the form of autunite (Krawczyk-Bärsch et al., 2015). A series of interesting responses with respect to prolonged uranium exposure in the filamentous cyanobacterium, A. torulosa, was observed leading to uranium precipitation and
ACCEPTED MANUSCRIPT biomass regeneration (Fig. 3) (Acharya et al., 2017). Cells of A. torulosa accumulated uranium in polyphosphate bodies within 24 h of U (100 µM) exposure under phosphate limited conditions. Further incubation of the cells (48 h) caused cell lysis and hydrolysis of polyphosphates with concomitant release of U and Pi. Cell bleaching, akinete formation followed
T
by uranium precipitation were observed following 120 h of uranyl exposure in this organism.
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The precipitated U (in the form of autunite mineral) settled at the bottom of the reaction vessel
CR
and provided a U “free” environment. Subsequently, inducible alkaline phosphatase activity in the akinetes led to germination of akinetes and complete regeneration of A. torulosa culture by
US
384 h (16 d) of U exposure. This phenomenon presented unusual insights into various metabolic
AN
responses adopted by A. torulosa to resist and survive under sustained uranium toxicity (Acharya et al., 2017).
Such responses can be utilized to mitigate uranium contamination for prolonged
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period.
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3.2. Bioprecipitation of uranium by yeast and fungal species Bioprecipitation of uranium has been observed in several yeasts and fungal species. In S. formation
of HUO 2 PO 4 ·4H2 O
PT
cerevisiae,
(H-autunite) was observed
because of the
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complexation of adsorbed uranium with the phosphate released from the cells. The mineral was formed at the ruptured regions of cell surfaces due to the localized saturation of the released
AC
phosphate (Ohnuki et al., 2005). Several other yeast species such as Kluyveromyces lactis, Pichia acacia, Cryptococcus podzolicus, Cryptococcus filicatus, Candida sake and Candida argentea have also been studied with respect to uranium bioprecipitation (Liang et al., 2016). Uranium phosphate precipitates were observed on the yeast cell surface on addition of organic phosphate substrates such as G2P or phytic acid in the medium. The precipitated uranium complexes were identified
as
chernikovite,
meta-ankoleite
K 2 (UO 2 )2 (PO 4 )2 ·6(H2 O),
bassetite
ACCEPTED MANUSCRIPT Fe++(UO 2 )2 (PO 4 )2 ·8(H2 O) and uramphite (NH4 )(UO 2 )(PO 4 )·3(H2 O) (Liang et al., 2016). Filamentous fungi like Aspergillus niger and Paecilomyces javanicus precipitated uranium significantly despite of their reduced growth in uranium containing media. The hyphal matrix of fungi served as the site for localization of uranyl phosphate minerals (Liang et al., 2015).
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4. Bioreduction of uranium
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Microbial reduction of metals involves the immobilization of potentially toxic soluble
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metals and radionuclides in their insoluble form by alteration of their oxidation states. Early reports on uranium reduction focused on abiological reactions with reductants such as sulfides,
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molecular hydrogen (H2 ) and organic compounds (Maynard, 1983). However, it is now evident
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that some bacteria can mediate reduction of the soluble oxidized form of uranium U(VI) to the insoluble tetravalent form, U(IV) (Wall and Krumholz, 2006; Newsome et al., 2014).
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Bioreduction is mostly observed in anaerobic bacteria which grow using uranium as potential
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electron acceptor or in other words, these microbes are said to respire uranium (Majumder and Wall, 2017). Since uranium bioreduction results in the formation of insoluble mineral uraninite,
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UO2 , it is also interchangeably termed as biomineralization. Uranium reduction in various
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microbes is detailed in this section (Fig. 1). 4.1. Bioreduction of uranium by Gram positive obligate anaerobic bacteria
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Several species of Clostridia have the inherent property of reducing U(VI) to U(IV) under acidic conditions (pH 5-6) (Gao and Francis, 2008). Clostridium is proposed to be one of the most important organisms for effectuating uranium reduction under natural conditions such as uranium pit water (Petrie et al., 2003), military facilities (Dong et al., 2006) and nuclear material storage sites (Madden et al., 2007). Fermentative processes in Clostridia are associated with U reduction (Petrie et al., 2003).
ACCEPTED MANUSCRIPT A spore forming sulfate-reducing bacterium, Desulfotomaculum reducens MI-1 isolated from metal contaminated sediments could grow by using U(VI) as the terminal electron acceptor in the absence of sulfate (Tebo and Obraztsova,1998). This organism could couple the oxidation of organic compounds with reduction of U(VI) to U(IV) and other metals such as manganese
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[Mn(IV) to Mn(II)], iron [Fe(III) to Fe(II)] or chromium [Cr(VI) to Cr(III)] for its growth.
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Spores of this bacterium could also bring about bioreduction of uranium by using H2 as the
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electron donor (Junier et al., 2009).
4.2. Bioreduction of uranium by Gram negative facultative anaerobic bacteria
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Mechanisms for uranium reduction in bacteria are not fully understood. Membrane
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associated electron transport respiratory (ETR) systems containing c-type cytochromes play an important role in uranium reduction (Lloyd, 2003). U reducing facultative anaerobic bacteria
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prevents the formation of sulfides and hydrogen sulfide complexes that have toxic and inhibitory
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effects on cellular metabolism (Mtimunye and Chirwa, 2014). In such microorganisms, extracellular U(VI) reduction is observed which involves electron flow through NADH-
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dehydrogenase, a primary electron donor associated with the ETR system. Salmonella
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subterranean (an acid tolerant bacterium and an important component of U(VI) reducing enrichment culture isolated from nitrate and uranium contaminated subsurface sediments)
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demonstrated significant reduction of uranium at pH 4.5 (Shelobolina et al., 2004). Members of the genus Shewanella have been extensively studied with respect to bioreduction of uranium. Initial studies on Shewanella putrefaciens demonstrated that the bioreduction of uranium was facilitated by enzymatic reactions coupled to electron transport chains within the organism (Gorby and Lovley, 1992). Shewanella alga, an iron reducing bacterium brought about reduction of uranium when complexed with multidentate aliphatic
ACCEPTED MANUSCRIPT ligands (malonate, oxalate and citrate) rather than monodentate acetate groups (Ganesh et al., 1997). This bacterium was unable to reduce uranium complexed with aromatic ligand and metal chelator such as tiron (4,5-dihydroxy-1,3-benzene disulfonic acid). Another important factor affecting uranium bioreduction kinetics is the bioavailability of uranium to the bacteria. U(VI)
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reduction by Shewanella oneidensis MR-1 in presence of sodium bicarbonate (NaHCO 3 ) showed
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slower reduction kinetics. The adsorption of U onto bacteria at pH 7.0 decreased with an increase
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in the concentration of NaHCO 3 due to the formation of soluble, negatively charged uranyl carbonate species (Sheng et al., 2011). A series of biological and chemical redox transformations
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in S. oneidensis including absorption of U(VI) to cell surface, formation of U(VI) nanowires
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[meta-schoepite (UO 3 .2H2 O)] and subsequent reduction to U(IV) nanoparticles [uraninite (UO 2 )] presented a novel strategy for U bioremediation in contaminated aquifers (Jiang et al., 2011).
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The two step uranium reduction was superior to the previously reported one step reduction
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process (Lovley et al., 1991). Mixed bacterial cultures derived from natural sediments and pure cultures of Desulfosporosinus species, Geobacter sulfurreducens, S. oneidensis MR-1 and S.
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putrefaciens CN32 are also reported to form biogenic uraninite nanoparticles (Suzuki et al.,
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2002; Renshaw et al., 2005; Senko et al., 2007; Bargar et al., 2008). In S. oneidensis MR-1, ctype cytochromes played an important role in reducing U(VI) to extracellular UO 2 nanoparticles
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(Marshall et al., 2006). The outer membrane (OM) associated decaheme cytochrome MtrC was essential for transferring electrons to U(VI) thereby reducing it to U(IV). Mutants of S. oneidensis MR-1 lacking mtrC (coding for cytochrome MtrC) or omcA (encoding for OM decaheme c-type cytochromes) displayed slower uranium reduction compared to the wild type strains (Marshall et al., 2006). Similarly, endecaheme c-type cytochrome (UndAHRCR-6) present in the OM of Shewanella sp. were involved in the extracellular reduction of iron [Fe(III)]
ACCEPTED MANUSCRIPT oxides and uranium [U(VI)] (Shi et al., 2011). Biofilms of Shewanella sp. HRCR-1 were also able to reduce uranium efficiently with the help of outer membrane c-type cytochromes present in the extracellular polymeric substances (Cao et al., 2011). However, S. oneidensis MR-1 and its biofilm exhibited limitations in reduction of U(VI) in flow and batch reactors (Sani et al., 2008).
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Apart from genetic manipulations, alternative organic materials/modules have been shown to
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facilitate and enhance uranium bioreduction rates. Anthraquinone-2,6-disulfonate provided as
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humus substitute improved the uranium reduction rates in S. oneidensis (Liu et al., 2015).
reduction rate of U(VI) (Suzuki et al., 2010).
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Similarly, FMN (Flavin mononucleotide) secreted by Shewanella species accelerated the
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4.3. Bioreduction of uranium by Gram negative obligate anaerobic bacteria In this section, the potential of Gram negative bacteria Geobacter metallireducens
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(previously designated as strain GS-15) and Anaeromyxobacter dehalogenans along with variety
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of sulfate reducing bacteria (SRB) in bringing about uranium bioreduction has been discussed. Geobacter species can mediate coupled reduction of Fe(III) and U(VI) in anaerobic subsurface
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environments. G. metallireducens, a Fe(III) reducing bacterium, showed direct reduction of
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U(VI) in contrast to indirect reduction of Fe(III). The reduction process promoted bacterial growth and the culture ceased to grow when uranium was depleted from the growth medium
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(Lovley et al., 1991). This organism reduced uranium in groundwater samples and the deposits of uraninite were observed on cell surfaces (Gorby and Lovely, 1992). Nitrate, a co-contaminant associated with uranium has a deleterious effect on reduction processes (Riley and Zachara, 1992) and it was necessary to add sufficient quantities of acetate (as an electron donor) initially to reduce nitrate (Finneran et al., 2002). Reduction and precipitation of U on surfaces harboring abundant electron donors cause damage to the cell envelope and interferes with the associated
ACCEPTED MANUSCRIPT functions. In order to safeguard cellular integrity and viability, conductive pili in Geobacter sp. demonstrated extracellular reduction of U(VI) to U(IV) (Cologgi et al., 2011). Pili were identified as primary uranium reductases that accepted electrons from the cell envelope and ctype cytochromes (Fig. 1). These served as electrical conduits between the cells and uranium.
in turn,
reduced
cell viability and cellular respiration
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periplasmic mineralization which,
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The lack of pili (Pil A- mutant) impaired U reduction in Geobacter and caused enhancement of
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(Reguera, 2012). A sulfate reducing species of Geobacter namely, G. sulfurreducens initially reduced U(VI) to a U(V) intermediate which on disproportionation resulted in the formation of
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tetravalent uranium (Renshaw et al., 2005). Cytochromes associated with outer surfaces (c-type)
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played an important role in U reduction in this organism (Orellana et al., 2013). Moreover, periplasmic cytochrome MacA, a diheme c-type cytochrome peroxidase, was found to be
M
essential for U(VI) reduction in G. sulfurreducens and a fraction of the reduced uranium was
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localized in the periplasm (Shelobolina et al., 2007). Growth yields of Geobacter lovleyi and G. sulfurreducens were lowered when U(VI) was used as electron acceptor suggesting that this
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process of reduction imposed additional burden to the growing cells (Sanford et al., 2007).
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Geobacter species have played a significant role in bioremediation of uranium under microcosm and in situ conditions. Microcosm studies using soil samples from Oak Ridge, FRC,
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TN, U.S. showed the enrichment of the members of Geobacteraceae family that could reduce uranium under low bicarbonate concentrations (1 mM) in the presence of ethanol as the electron donor (Luo et al., 2007). Uranium bioreduction continued up to 50 days during a field trial at Rifle, CO, U.S. (Anderson et al., 2003). With acetate supplementation (as an electron donor), concentrations of U(VI) were observed to decrease from 0.4 to those below 0.18 μM which was considered to be the maximum contaminant limit. In situ reduction of U(VI) in this site was not
ACCEPTED MANUSCRIPT effective at greater depths (4.0 to 5.2 m) as the substantial proportion of available uranium was not amenable for bioreduction (Ortiz-Bernad et al., 2004). The microbial community at a uranium contaminated field site in Rifle Integrated Field Research Centre (IFRC), CO, U.S. was also dominated by Geobacter species when biostimulated with acetate. These bacteria effectively
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brought about dissimilatory uranium reduction (Mouser et al., 2009).
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An anaerobic myxobacterium isolate obtained from Oak Ridge FRC, TN, U.S. was
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identified as A. dehalogenans. This culture reduced uranium by using H2 as the electron donor in contrast to acetate that was utilized by Geobacter (Wu et al., 2006). While alternative electron
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acceptors such as Fe(III) citrate or Fe(III) oxide inhibited U(VI) reduction in A. dehalogenans,
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fumarate and 2-chlorophenol did not show such inhibitory effect (Wu et al., 2006). The SRB, Desulfovibrio desulfuricans, demonstrated rapid and direct reduction of uranium in the presence
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of lactate or H2 supplied as electron donors (Lovley and Phillips, 1992a). U(VI) and sulfate
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reduction occurred simultaneously and reduced uranium was observed in the form of extracellular uraninite. D. desulfuricans cells contained within a semipermeable membrane
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showed rapid reduction of uranium at concentrations as high as 24 mM (Lovley and Phillips,
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1992b). Bicarbonate extraction of uranium from contaminated soils and subsequent microbial reduction by D. desulfuricans demonstrated a potential strategy for concentrating U from
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contaminated soil and sediments (Phillips et al., 1995). Cytochrome c3 appeared to be integral for in vivo electron pathway that was involved in the reduction of U(VI) to U(IV) in D. desulfuricans. A cytochrome c3 mutant strain of D. desulfuricans G20 was less effective (almost by 50%) in uranium reduction than wild type strains in presence of lactate or pyruvate as electron donors (Payne et al., 2002). A variety of sulfate reducing bacteria (SRB) were analyzed for their ability to reduce uranium (Lovley et al., 1993a). Among different isolates, Desulfovibrio vulgaris
ACCEPTED MANUSCRIPT was the most effective uranium reducing organism followed by Desulfovibrio baculatus, Desulfovibrio sulfodismutans and Desulfovibrio baarsii (Lovley et al., 1993a). Uranium reduction by all these SRBs was highly dependent on the supplementation of electron donors. D. vulgaris was unable to grow in presence of U(VI) as the sole electron acceptor, precipitated
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uranium as uraninite and reduced sulfate and uranium simultaneously (Lovley et al., 1993b)
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similar to D. desulfuricans (Lovley and Phillips, 1992a). Another SRB, Desulfosporosinus sp.
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(isolate P3) and its type strain Desulfosporosinus orientis (DSM 765) showed enzymatic reduction of uranium using lactate or H2 as electron donor (Suzuki et al., 2004).
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A field study related to growth and activity of indigenous microorganisms in the presence
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of acetate, glucose or ethanol was carried out at Oak Ridge, FRC, TN, U.S. (Istok et al., 2004). Supplementation with these electron donors resulted in the development of an anaerobic and
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reducing environment that favored reduction of NO 3 -, Fe(III) and U(VI). NO 3 - dependent,
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microbially mediated U(IV) oxidation is an important process in modulating the stability of bioreduced U(VI) at sites contaminated with high concentration of nitrates. In situ uranium
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bioremediation trials with acetate supplementation were carried out at Rifle, IFRC, CO, U.S.
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Rapid reduction of uranium prevalent in the form of Ca-UO 2 -CO3 ternary complexes was achieved by iron and uranyl respiring Geobacter species (Williams et al., 2011). In a later study,
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at the same site, U reduction was observed in presence of acetate, lactate, hydrogen release compound (HRC) or vegetable oil. The composition of microbial communities varied with different donors (Barlett et al., 2012). Another report at Oak Ridge IFRC, TN, U.S. describes U(VI) bioreduction over an extended (one year) period. Emulsified vegetable oil (EVO) amendments caused effective bioreduction at depths up to 50 m (Watson et al., 2013). Recent studies at Oak Ridge, TN, U.S. have demonstrated U(VI) bioreduction in iron oxide-rich
ACCEPTED MANUSCRIPT sediments after ethanol supplementation (Li et al., 2018). Major problem associated with in situ bioremediation of uranium contaminated sites is the fate of reduced U(IV) in sub-surfaces which re-oxidizes to U(VI) form and affects the process economy negatively. A recent review focuses on the challenges and complexities associated with remediation of uranium contaminated soils
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(Selvakumar et al., 2018). Factors such as soil type, presence of organic matter, co-ions along
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with soil pH and uranium species play a major role in determining the efficiency of
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bioremediation methods and should be considered while designing strategies for uranium removal from contaminated soil.
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5. Uranium bioremediation by genetically engineered bacteria
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The use of genetically engineered microorganisms (GEMs) for bioremediation of metal polluted environments has become important in the recent years. There are several reports
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highlighting the potential of GEMs for removing toxic metals and radionuclides (Garbisu and
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Alkorta, 1999; Urgun-Demirtas et al., 2006; Menn et al., 2008; Kulkarni et al., 2013). Relevant genes have been effectively expressed in appropriate hosts and the resulting GEMs have
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demonstrated remarkable bioremediation efficiencies (Appukuttan et al., 2006, 2011; Nilgiriwala
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et al., 2008; Kulkarni et al., 2013, 2016). Majority of the studies utilizing genetically engineered microorganisms have focused on
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phosphatase mediated uranium precipitation rather than on biosorption or bioreduction. For example, Escherichia coli cloned with a non-specific acid phosphates gene, phoN showed superior (2.5 times) uranium precipitation than Citrobacter sp. constitutively expressing the phoN gene (Basnakova et al., 1998). The uranyl precipitates in the recombinant E. coli cultures were identified as HUO 2 PO4 . The radiosensitivity of E. coli, Citrobacter sp. or Pseudomonas sp. seemed to limit their usage in treatment of radioactive waste. Recombinant bacteria expressing
ACCEPTED MANUSCRIPT phosphatases and capable of precipitating uranium under acid or alkaline conditions even after being exposed to high doses of ionizing radiation have been reported (Appukuttan et al., 2006; Appukuttan et al., 2011; Kulkarni et al., 2013). Deinococcus radiodurans is known to tolerate extremely high doses of ionizing radiations well above those encountered under natural
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conditions (Battista, 1997). Recombinant cells of D. radiodurans expressing a non-specific acid
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phosphatase, PhoN (DrPhoN) were shown to precipitate uranium efficiently in presence of G2P
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under acidic conditions (Appukuttan et al., 2006). However, recombinant cells of D. radiodurans exhibited slower uranium precipitation (70% after 26 h) as compared to E. coli cells expressing
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PhoN (70% in 2 h) when exposed to aqueous solutions containing 0.8 mM uranyl nitrate at pH
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5.0 (Appukuttan et al., 2006). The multilayered cell envelope of D. radiodurans apparently decreased the accessibility of the organophosphate substrate to the periplasmic PhoN as
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compared to E. coli. Recombinant D. radiodurans cells retained 90% of their uranium 60
Co gamma rays
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precipitation abilities even after exposure to very high doses (6 kGy) of
presenting their potential for treating radioactive nuclear waste. Radiation sensitive E. coli
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clones, on the other hand, lost their ability to precipitate U following exposure to lower doses (1
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kGy) of gamma rays (Appukuttan et al., 2006). Moreover, the lyophilized DrPhoN cells showed appreciable loading of bioprecipitated uranium up to 5.7 g U g-1 dry weight. This property was
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retained even after 6 months of storage at room temperature. Fresh (non-lyophilized) cells on the other hand, lost their ability to precipitate U after 8-10 days of storage at 4˚C (Appukuttan et al., 2011) establishing the enhancement of shelf-life of lyophilized cells for application purposes. It was observed that the recombinant D. radiodurans cells expressing PhoN from a radiation inducible ‘ssb’ promoter exhibited enhanced phosphatase activity following exposure to gamma
ACCEPTED MANUSCRIPT radiation and removed uranium more rapidly as compared to the cells expressing PhoN from a radiation non-inducible Deinococcal groESL promoter (Misra et al., 2014). Uranium bioprecipitation from alkaline solutions has also been attempted by using recombinant bacteria. Recombinant E. coli or D. radiodurans cells expressing Sphingomonas
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PhoK (alkaline phosphatase) demonstrated high enzyme activity and precipitated U from
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alkaline solutions (pH 9.0) very effectively (Nilgiriwala et al., 2008; Kulkarni et al., 2013;
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Kulkarni et al., 2016). Recombinant D. radiodurans strain expressing PhoK revealed very high loading (10.7 g U g-1 dry weight of cells) of uranium (Kulkarni et al., 2013). These cells 60
Co gamma rays. When
in calcium alginate, they exhibited superior uranium removal from alkaline
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immobilized
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precipitated uranium effectively even after exposure to 6 kGy dose of
solutions (Kulkarni et al., 2013).
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There are few reports on development of recombinant bacterial strains for uranium Asn
128T yr change in the chromate
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bioreduction and bioaccumulation. A single amino acid
reducing enzyme ChrR6 from E. coli resulted in significant improvement in the kinetics of
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uranyl reduction and minimized the chances of reoxidation and resolubilization of the reduced
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product (Barak et al., 2006).
Metallothioneins (MT) are small, cysteine-rich, intracellular, metal binding proteins that
E.
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are implicated in metal detoxification (Pettersson et al., 1988; Daniels et al., 1998). Recombinant coli cells expressing cyanobacterial (Synechococcus) metallothionein, SmtA exhibited
increased
tolerance towards uranium (Acharya and
Blindauer, 2016). However, SmtA
homologue, namely NmtA from another cyanobacterium, Anabaena sp. strain PCC 7120 when expressed in E.coli did not confer any notable uranium tolerance (Divya et al., 2018). The recombinant SmtA was able to bind uranium in the presence of excess carbonate ions under in
ACCEPTED MANUSCRIPT vitro conditions via its glutamate and aspartate residues without compromising its secondary or tertiary structure (Acharya and Blindauer, 2016). The uranyl binding affinity of recombinant SmtA, however, was lesser (KD = 100 pM) than that of a protein engineered for high uranyl binding, namely, super uranyl binding protein (SUP) (KD = 7.4 fM). SUP was obtained by
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multiple mutations of the ligand residues (to aspartate/asparagine or glutamate/glutamine) in a
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small protein of unknown function in Methanobacterium thermoautotrophicus (Zhou et al.,
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2014). SUP when fused with maltose binding protein could remove 90% of U from synthetic sea water in presence of all major ions. SUP was displayed on E. coli surface with OmpA (outer
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membrane protein) as a fusion protein and the resulting recombinant E. coli cells could sequester
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60% of U from synthetic sea water (Zhou et al., 2014). Similarly, a uranyl responsive DNA binding protein was designed by engineering the nickel (IIs) responsive protein, NikR from E.
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coli. The resulting recombinant NikR protein bound uranyl with a dissociation constant K D= 53
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nM and selectively bound to DNA in presence of uranium (Wegner et al., 2009). Although genetically engineered microorganisms display potential interactions with U,
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their exploitation for bioremediation applications requires further validation.
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6. Understanding uranium interactions in microbes via ‘omics’ based approaches Recent advances in analytical and bioinformatics techniques have made it possible to
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understand gene/protein expression under different stress conditions via genomic and proteomic approaches. Genomics and metagenomics are important in establishing genetic variability, population structure and ecological functions of microorganisms (Riesenfeld et al., 2004; Xu, 2006). We discuss here, genomics, transcriptomics and proteomics studies in microbes in response to the uranium exposure (Fig. 1). 6.1. Genomics, transcriptomics and metagenomics investigations in the presence of uranium
ACCEPTED MANUSCRIPT Such studies allow assessment of the impact of metals on microbial communities (Akob et al., 2006; Suriya et al., 2017). Microorganisms living in contaminated sites exhibit different strategies for counteracting toxic effects of metals. For example, the occurrence of P IB-type ATPases (metal transporters) was established in uranium tolerant, aerobic heterotrophic bacteria
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isolated from unmined uranium rich deposits in North-East India as a result of horizontal gene
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transfer (Nongkhlaw et al., 2012).
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Gene expression studies have helped in understanding the mechanism of uranium tolerance in microorganisms. Transcriptome profiling of Geobacter uraniireducens and S.
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oneidensis MR-1 has illustrated a varied stress response in the presence of uranium (Bencheikh-
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Latmani et al., 2005; Holmes et al., 2009). Transposon mutagenesis in C. crescentus revealed the importance of outer membrane transporters (rsaFa and rsaFb ), a stress-responsive transcription
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factor (cztR), or a ppGpp synthetase/hydrolase (spoT) in conferring uranium tolerance. RsaF a and
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RsaFb were shown to efflux U and maintain the membrane integrity during U stress (Yung et al., 2015). Transcriptomics studies on Metallosphaera prunae (a thermoacidophilic bacterium
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isolated from a uranium mine in Thuringen, Germany) exposed to high concentrations of soluble
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uranium (1,238 mg l-1 ) were carried out (Mukherjee et al., 2012). Substantial degradation of cellular RNA and abortion of transcriptional and translational processes were observed in the
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organism within 15 min of U exposure possibly for resisting uranium toxicity. However, RNA integrity was restored within 60 min of uranium treatment. Possible role of siderophores in U sequestration in M. prunae was suggested as the genes related to iron complex transport system were found to be significantly induced during U exposure (Mukherjee et al., 2012). Metagenomic analysis of microbial community from uranium contaminated sub surfaces of Oak Ridge FRC, TN, U.S., Ervas Tenras Pinhel Guarda, Portugal, Cauvery River, Tamil
ACCEPTED MANUSCRIPT Nadu,
India
and
uranium ore
deposit
of Domiasiat,
India,
showed
abundance
of
Gammaproteobacteria (Akob et al., 2006; Caetano et al. 2014; Suriya et al., 2017; Kumar et al., 2013b). Similar observations were made with respect to uranium polluted sites at Bessines, Limousin, France (Mondani et al., 2011) and uranium mines at Jaduguda, Bagjata and Turamdih
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located in East Singhbhum district, Jharkhand, India (Dhal et al., 2011). Proteobacteria like A.
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dehalogenans and Geobacter were found in abundance at Oak Ridge National Laboratory, TN,
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U.S. (North et al., 2004) and the dominance of Geobacter species like Geobacter bemidjiensis, G. metallireducens, G. sulfurreducens PCA, G. lovleyi SZ, G. uraniireducens Rf4, Geobacter
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strain FRC-32 and Geobacter strain M21 was observed at U contaminated sites in Rifle, CO,
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U.S. (Wilkins et al., 2009). Members of Rhodocyclaceae family of Proteobacteria group were found in samples collected from different locations of U contaminated sites of Portugal (Martins
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et al., 2010). Burkholderia sp. strain SRS-W-2-2016 (Proteobacteria group) isolated from
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Savannah River Site, U.S. displayed ecologically relevant genomic traits such as substrate binding proteins, permeases, transport regulators and efflux pumps to restrict metals in the
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extracellular environment or prevent their cellular uptake. This aided survival of the bacterium in
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the presence of mixed contaminants (Pathak et al., 2017). Apart from Proteobacteria, Actinobacteria and Firmicutes have also been reported from
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Oak Ridge, FRC, TN, U.S. and uranium mine in Jaduguda, India (Bollmann et al., 2010; Islam and Sar, 2011). Similarly, bacterial community of soil samples collected from T22 trench of the Chernobyl, Ukraine, was dominated by Firmicutes (Chapon et al., 2012). Desulfotomaculum like microorganisms (Firmicutes) were dominant in uranium mill tailings disposal site at Shiprock, New Mexico (Chang et al., 2001). Proteobacteria and Acidobacteria were reported from soil samples of uranium ore deposit of Domiasiat in Northeast India (Kumar et al., 2013b).
ACCEPTED MANUSCRIPT Metagenomic analysis of a radioactive contaminated soil at uranium tailings from southern China showed structural and functional diversity of the soil bacterial community with abundance of Actinobacteria and Proteobacteria groups and proteins related to membrane transport and carbohydrate metabolism (Yan et al., 2016).
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A recent study on microbial community analysis of uranium contaminated aquifers at
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Oak ridge IFRC, TN, U.S. has shown the dominance of SRB such as Desulfococcus,
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Desulfobacterium and Desulfovibrio (Zhang et al., 2017). The functional diversity of microorganisms was shown to decrease in the groundwater samples at higher uranium
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concentrations (He et al., 2018). The levels of uranium contamination and ecosystem functioning
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could be predicted on the basis of such variations in microbial diversity. The effect of varying uranium concentrations (0-4000 mg kg-1 ) on microbial diversity was examined in Gulungul
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billabong near Ranger uranium mines, Australia (Sutcliffe et al., 2017a). Archaeal and bacterial
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components prevalent in methanogenic consortia were observed at higher concentrations. Members belonging to Orders Rhizobiales and Acidobacterial were inhibited by uranium. The
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genes associated with methanogenesis, nitrogen fixation and metal efflux showed a positive
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correlation with uranium concentration while the genes involved in aerobic respiration and amino acid transporters were negatively affected (Sutcliffe et al., 2017a). Another study on
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microbial communities associated
with sediments of the same site with high uranium
concentrations (4000 mg kg-1 ) showed the predominance of five bacterial genera namely, Geobacter, Geothrix, Dyella, Bacteroidetes and Chloroflexi (Sutcliffe et al., 2017b). Overall, it is evident from the cited studies that members of Phylum Proteobacteria are most dominant members in uranium contaminated environments. 6.2. Proteomic responses in presence of uranium
ACCEPTED MANUSCRIPT Proteomic studies have been carried out to understand the mechanisms involved in uranium tolerance and detoxification. In this section, comparative proteomic analysis in response to uranium observed in different groups of microorganisms has been discussed. The major alterations
in
protein
profiles
following
uranium
exposure
in
some
representative
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6.2.1. Proteomic responses of bacteria in presence of uranium
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microorganisms are highlighted in Table 3.
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Proteome analysis of C. crescentus exposed to U revealed down regulation of cell cycle regulators, motility, chemotaxis proteins and up regulation of a putative phytase (Yung et al.,
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2014). Phytase provided the requisite phosphate groups for uranium precipitation and favored
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cell survival of C. crescentus in presence of U (Yung and Jiao, 2014). Increased expression of proteins involved in general stress response or in detoxification of reactive oxygen species was
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observed in U exposed cells of A. ferrooxidans (Dekker et al., 2016). Similarly, E. coli cells
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exposed to uranyl ions showed differential accumulation of oxidative stress proteins and other proteins like NADH/quinone oxidoreductase WrbA (Khemiri et al., 2014). Low molecular
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weight thiols played an important role in protecting cells of G. sulfurreducens, E. coli and A.
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ferrooxidans against U induced oxidative stress (Orellana et al., 2014; Khemiri et al., 2014; Dekker et al., 2016). G. sulfurreducens displayed increased levels of several efflux pumps
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belonging to the RND (resistance-nodulation-cell division) family, superoxide dismutase and superoxide reductase when exposed to U (Orellana et al., 2014). Similarly, the higher tolerance of uranium in Anabaena strain L-31(LD50 of 200 µM) as compared to Anabaena strain PCC 7120 (LD50 of 75 µM) was attributed to the increased levels of oxidative stress defense protein like Manganese SOD (MnSOD) in L-31 as compared to PCC 7120 (Panda et al., 2017). Proteins involved in phosphate and iron metabolism were found to be abundant in a uranium tolerant
ACCEPTED MANUSCRIPT bacterium, Microbacterium oleivorans A9, isolated from a sample near Chernobyl nuclear power plant. The role of siderophores was proposed in uranium transportation in this bacterium (Table 3) (Gallois et al., 2018). 6.2.2. Uranium induced proteomic responses in yeasts U(VI) or
U(VI) and increased expression of specific uncharacterized proteins following U exposure
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233
238
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S. cerevisiae strain BY4743 showed reduced growth when exposed to
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(Sakamoto et al., 2007). Deletion of the phosphate transporter genes of PHO86, PHO84, PHO2, and PHO87 resulted in the uranium-sensitive (mutant) yeast strains suggesting that the phosphate
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transporting genes contributed to uranium tolerance. Deletion of genes encoding for cell
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membrane proteins, phospholipid-binding proteins and cell wall proteins in some of the uncharacterized yeast strains demonstrated lower uranium accumulation in the mutant strains
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indicating the positive role of the cell surface proteins in the accumulation process (Sakamoto et
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al., 2012).
6.2.3. Metaproteomics analysis of uranium exposed communities
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Metaproteomics studies of environmental samples have become possible with recent
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advances in development of protein extraction methods from groundwater, soils and sediments (Wrighton et al., 2012; Handley et al., 2013). The metaproteome profiles revealed changes in the
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microbial activities and identified active biochemical pathways and functions specific to microbial groups after biostimulation withemulsified
vegetable oil (EVO) at a uranium
contaminated site (Chourey et al., 2013) Members of Betaproteobacteria (Dechloromonas, Ralstonia, Rhodoferax, Polaromonas, Delftia and Chromobacterium) and Firmicutes and the proteins of P. putida, Geobacter species and Dechloromonas like microbes were found to be abundant in contaminated sites. Moreover, proteins responsible for ammonium assimilation,
ACCEPTED MANUSCRIPT EVO
degradation
and
polyhydroxybutyrate granule formation were prominent following
biostimulation (Chourey et al., 2013). A proteogenomic study during in situ U bioreduction by Geobacter members at Oak Ridge FRC, TN, U.S. revealed that the proteome of these strains were dominated by enzymes which converted acetate to acetyl-CoA and pyruvate throughout the
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biostimulation process (Wilkins et al., 2011). The identification of functional genes and proteins
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implicated in metal responses are fundamental in deciphering underlying molecular mechanisms
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and for developing in situ bioremediation strategies. 7. Conclusions
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In the recent times, uranium contamination resulting from anthropogenic activities has
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affected the microbes in various ways. Due to lack of clarity on the biological relevance of uranium, microbial interactions with uranium have been explored less as compared to other
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potentially toxic metals including arsenic, cadmium, mercury, silver and chromium. It is
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therefore important to understand microbial machinery in response to U to develop in situ remediation measures for controlling uranium contamination. A summary of the interactions of
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microorganisms with uranium along with ‘omics’ tools used for understanding these interactions
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are depicted in Fig. 1. The use of microbial systems is advantageous for providing cost effective solutions without the generation ofhazardous secondary wastes. Scale-up procedures and field
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trials have been undertaken on many occasions for uranium bioremediation. Recent advances in genomics, transcriptomics, metagenomics, proteomics and metaproteomics have helped us in understanding the significance of genes,
proteins or microbial communities capable of
interacting with uranium. In the future, the natural U tolerant environmental microbes as well as the microbial metabolic products obtained using ‘omics’ based approaches could assist in natural attenuation of U contamination.
ACCEPTED MANUSCRIPT Conflicts of interest The authors declare that they have no competing interests . Acknowledgments SZ and NK thank University Grants Commission for funds under University for Potential
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Excellence (UPE) Phase II. NK thanks SPPU-BARC collaborative Ph.D. Programme for
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financial support. We gratefully acknowledge the help of Dr. Sharda Sawant and Ms. Siddhi
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Redkar, Advanced Centre for Treatment, Research and Education in Cancer (ACTREC), Mumbai for TEM analysis and Microscopy Division, Carl Zeiss, India (Bangalore), Zeiss Group,
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for SEM analysis of the yeast samples.
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Table 1. Details of microorganis ms displaying high tolerance towards uranium Microorganism Tolerance/Minimum Site of isolation inhibitory concentration (MIC) (mM) Candida sorbophila 13 Uranium mine, Salamanca, Spain Acidithiobacillus 9 Uranium waste piles, ferrooxidans Johanngeorgenstadt, Germany Pseudomonas aeruginosa 5 Uranium ore deposit, DPs-13 Domiasiat, India Thermoterrabacterium 5 Hot springs, Yellow ferrireducens stone National Park, WY, U.S. Desulfovibrio desulfuricans 4 Wealden clay, G20 and I2 England Microbacterium oxydans 4 Radioactive waste depository, Russia Serratia marcescens 4 U ore deposit, Domiasiat, India Shewanella putrefaciens 200 3 Crude oil pipeline, Canada 3
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Rhodotorula mucilaginosa
Uranium mine, Salamanca, Spain
Reference
de Silóniz et al. (2002) Merroun and SelenskaPobell (2001) Sarma et al. (2014) Khijniak et al. (2005) Payne et al. (2002) Nedelkova et al. (2007) Kumar et al. (2011) Wade and DiChristina (2000) de Silóniz et al. (2002)
ACCEPTED MANUSCRIPT Table 2. Uranium biosorption capacities of different Microorga Bio Uraniu p Temper Equilib nism mass m H ature rium quan concent (ºC) time tity ration (min (mg) (mM) where not specifi ed) Halobacte 0.04 6. rium 0 noricense DSM15987 Geobacill 250 4.20 5. us 0 thermoleo vorans subsp strombolie nsis Bacillus 150 0.0042- 4. mojavensi 0.021 0s 7. 0
microorganisms Isotherm, Chemical Kinetic groups model, and/or Thermod molecules ynamic for U parameter binding s Phosphoryl, carboxyl
9.3
Bader et al. (2017 )
-
11
Özde mir and Kilinc (2012 )
-
-
48.2
-
-
-
50
Özde mir et al. (2017 a) Özde mir et al. (2017 b) Tsuru ta (2004 ) Li et al. (2016 )
4. 05. 0
-
-
60
-
-
9.04
3. 5
100
0.042
3. 0
Room temper ature
12 h
Freundlic h, pseudo second order
Amino, phosphate, carboxyl
>3.0
-
0.0289
7. 3
25
24 h
Langmuir , pseudo second order
-
84
0.1
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Streptomy 15 ces levoris
Streptomy ces sporoverr ucosus dwc-3 Arthrobac ter G975
0.00420.021
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100
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Bacillus vallismort is
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Load Refer ing ence (mg g-1 )
Carva jal et al. (2012 )
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Amycolato 0.00 psis sp. 1 K47
0.168
4. 0
40
Equilib rium time (min where not specifi ed) 150
Brachyba cterium sp. G1
0.04
6. 0
-
-
Myxococc 20 us xanthus
1
4. 5
28
10
Citrobacte r freudii
0.0840.63
-
Chemical groups and/or molecules for U binding
Load Refer ing ence (mg g-1 )
Langmuir ΔH° = 22.50 kJ mol-1 ΔS°= 0.014 kJ mol-1 K -
Carboxyl, hydroxyl, amide
38.8
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p H
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Bio mass quan tity (mg)
-
Carboxyl
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20-50
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25-55
Pseudomo nas aeruginos
3050
0.4
2. 4
22
-
Langmuir Carboxyl and Freundlic h, pseudo second order, ΔG0 = 6.54 kJ mol-1 spontaneo us, ΔH0 = 12.69kJ mol-1 endother mic, ΔS0 = 64.53 J mol-1 K -
Celik et al. (2018 )
971 Bader ± 29 et al. (2018 ) 28.5 Gonz 6 alezMuno z et al. (1997 ) 48 Xie et al. (2008 )
100
Hu et al. (1996
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Temper ature (ºC)
Equilib rium time (min where not specifi ed)
Chemical groups and/or molecules for U binding
Load Refer ing ence (mg g-1 )
) 7. 8
-
60
0.1
7. 8
-
24 h
Microcysti s aeruginos a
0.168
7. 0
-
Saccharo myces cerevisiae (heat killed) Rhodotoru 4 la glutinis (raw)
0.0042 30 5. 5
Langmuir
Amide, carboxyl
53.5
Achar ya et al. (2009 ) Achar ya et al. (2012 ) Li et al. (2004 )
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0.1
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15
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Exopolysaccharide, Polyphosphate bodies
77.3 5
Langmuir , Freundlic h isotherm -
-
44
Hydroxyl, carboxyl, phosphate, amino
29.8
Lu et al. (2013 )
Langmuir and Freundlic h Langmuir , pseudo second order, ΔG0 = 21.9 ± 0.1 kJ mol-1 spontaneo us, ΔH0 = 36.1 ±1.9
Amino, carboxyl
98.4
-
26
Bai et al. (2010 ) Bai et al. (2012 )
-
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0.588 25 6. 0
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Rhodotoru 5.15 la glutinis (magnetic ally modified)
Isotherm, Kinetic model, Thermod ynamic parameter s
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a strain CSU Synechoco ccus elongatus strain BDU/750 42 Anabaena torulosa
Uraniu m concent ration (mM)
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Microorga nism
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0.420 20-30 6. 0
-
30
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Uraniu m concent ration (mM)
p H
Temper ature (ºC)
Equilib rium time (min where not specifi ed)
Isotherm, Kinetic model, Thermod ynamic parameter s
Chemical groups and/or molecules for U binding
kJ mol-1 endother mic, ΔS0 = 198.0± 6.3 J mol1 K irreversib le Sips Methyl, model, methylene pseudo first order, ΔG0 = 16.8 ± 0.1 kJ mol-1 spontaneo us, ΔH0 = 47.2± 3.7 kJ mol-1 endother mic, ΔS0 = 218 ± 12 J mol-1 K irreversib le Pseudo second order
Load Refer ing ence (mg g-1 )
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30
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0.153 25-6. 0
4.20
28 4. 5
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Trichoder 0.42 ma 0 harzianum
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Rhodotoru 5 la glutinis (immobili zed)
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Microorga nism
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Fusarium sp. #ZZF51
100
0.21
4. 0
Room Temp.
60
Lentinus sajor-caju MAFF 430306
-
0.840
4. 5
5-35
30
Langmuir , pseudo second order Freundlic h
17.3
Bai et al. (2014 )
196
Akhta r et al. (2007 ) Yang et al. (2012 ) Bayra moğlu et al.
Hydroxyl, carboxyl
16
Carboxyl, amino
128
ACCEPTED MANUSCRIPT Microorga nism
Bio mass quan tity (mg)
Uraniu m concent ration (mM)
p H
Temper ature (ºC)
Equilib rium time (min where not specifi ed)
Isotherm, Kinetic model, Thermod ynamic parameter s
Chemical groups and/or molecules for U binding
-
-
-
Padina sp. -
0.2104.20
4. 0
37
74
Carboxyl, sulfonate, hydroxyl
377
Cystoseria indica (calcium pretreated ) Cystoseria indica
1.47
4. 0
30
Langmuir , second order regressio n model Langmuir , pseudo second order
-
455
Khani et al. (2006 )
0.2104.20
4. 0
Langmuir , first order, Activatio n energy= -6.15 kJ mole -1 , ΔG0 = 8.7 kJ mole-1 spontaneo us, ΔH0 = -16.39 kJ mole-1 exothermi c, ΔS0 = 0.026 J mol-1 K reversible
-
198
Khani et al. (2008 )
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5. 06. 0
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Phosphate
(2006 )
0.00210.5
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300330
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Schizophy llum commune
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(Alkali treated)
Load Refer ing ence (mg g-1 )
120150 h
280
Gunth er et al. (2014 ) Khani (2011 )
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Temper ature (ºC)
Catenella repens
50
0.420
4. 5
15-45
Equilib rium time (min where not specifi ed) 45
Isotherm, Kinetic model, Thermod ynamic parameter s
Chemical groups and/or molecules for U binding
Load Refer ing ence (mg g-1 )
Langmuir , Freundlic h, pseudo second order
Carboxyl, hydroxyl, sulphate, phosphate
303
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Uraniu m concent ration (mM)
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Bhat et al. (2008 )
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Table 3. Summary of the proteomic responses observed in different microorganis ms after exposure to uranium Microorganis Uranium Exposu Protein levels upProtein levels Referen m Concentrati re time regulated down-regulated ce on (h) (µM) Escherichia 50 2 NADH/quinone Adenine Khemiri coli MG1655 oxidoreductase phosphoribosyl et al. (WrbA) transferase (Apt), (2014) polynucleotide Escherichia 80 2 phosphorylase coli MG1655 Fe-Superoxide (PNPase), chaperon Khemiri dismutase (SodB), ClpB, HtpG et al. monooxygenases, (Hsp90) (2014) WrbA Catalase (KatE), Apt Caulobacter 500 3 Two component cell Aerobic oxidative Yung et crescentus signaling, amino acid phosphorylation, al. CB15N metabolism, phytase, cell motility, (2014) ABC transporter, chemotaxis, cell nucleoid associated adhesion, Electron protein, Structural transport maintenance of chain(ETC), chromosomes protein folding, S(SMC) proteins, cell layer biosynthesis growth and division associated proteins Anabaena 75 3 Phycobilisome core Phycobilisome core Panda PCC 7120 component (ApcF), membrane linker et al. phycobilisome rod (ApcE-1 and (2017) core linker protein ApcE-2), Enolase (CpcG1, (Eno), Fructose 1,6 CpcG4),phycocyanobil bisphosphate in lyase CpcS2 aldolase (Fda-1), (CpcS2-1), ribuloseFe-superoxide 1,5-bisphosphate dismutase(FeSOD), carboxylase/oxygenase PNPase, molecular , cysteine synthase A chaperones GroEL (CysA-1), (chaperonin), peroxiredoxins (Prx), DnaK (heat shock + ferredoxin-NADP protein) reductase
ACCEPTED MANUSCRIPT Exposu re time (h)
Protein levels upregulated
Protein levels down-regulated
Referen ce
3
Pyruvate dehydrogenase E1 subunit (Pde1), peroxiredoxins (Prx), ferredoxinNADP+reductase(F NR)
Panda et al. (2017)
Geobacter species
-
-
Geobacter sulfurreducen s DL-1
100
4
ATP synthase FOF1 subunit alpha (AtpA), beta (AtpB-1, AtpB-2), DNA directed RNA polymerase (RpoB-1), Mn-superoxide dismutase (MnSOD), transketolase (Tkt-b), DNA repair protein (UvrB) Ribosomal proteins, ATP synthase F1 subunit, tricarboxylic acid (TCA) cycle proteins (citrate synthase and isocitrate dehydrogenase), acetate-activating enzyme acetylcoenzyme A (CoA) hydrolase RND (resistancenodulation-cell division) family, lipid bilayer of outer membrane for protection, secretion systems, chaperones, c-type cytochromes, superoxide dismutase and reductase
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Wilkins et al. (2009)
Central metabolism, amino acid biosynthesis (GSU1061, GSU3099, GSU3095 and GSU1828), translation (GSU 1920), ribosome biogenesis (ObgE and EngB) Outer membrane protein 40 (Omp40)
Orellan a et al. (2014)
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Anabaena L 31
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Uranium Concentrati on (µM) 200
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Microorganis m
Acidithiobacil lus ferrooxidans ATCC 23270
Microbacteri um oleivorans A9
500
-
10
24
Heat shock proteins(Hsps), cochaperone protein GrpE
Proteins involved in phosphate and iron metabolism, siderophores, DNA-
-
Dekker et al. (2016)
Gallois et al. (2018)
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Protein levels upregulated
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binding transcriptional (ArsR family), DNA binding response regulator (NarL/FixJ family), Transporters (lipoprotein, Fe3+hydroxamate, dipeptide/oligopeptide/ nickel, enterochelin, Fe3+/spermidine/putres cine)
Protein levels down-regulated
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Uranium Concentrati on (µM)
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Microorganis m
Referen ce
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Fig. 2: Electron microscopy analysis of U challenged Yarrowia lipolytica cells. SEM images of
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(A) U untreated control cells or (B) cells treated with 50 µM uranyl carbonate for 24 h at pH 7.5. U untreated cells show characteristic oval morphology and bud scars (white arrows) (A) whereas
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the cells treated with uranium exhibited highly irregular and crustaceous surface morphology (B). TEM images of (C) U untreated cells or (D) 24 h U treated cells. Intracellular (black arrows)
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and extracellular (white arrows) uranyl deposits were observed in U treated cells.
Fig. 3: Illustration of the metabolic events exhibited by a filamentous cyanobacterium,
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Anabaena torulosa following exposure to 100 µM uranyl carbonate under phosphate limited
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conditions at pH 7.8 for 384 h (16 days). Bright field images were acquired by using Carl Zeiss Axioscop 40 microscope with oil immersion objectives (magnification x1,500, bars indicate 5
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µm). Red arrows indicate uranium association with polyphosphate bodies (24 h) and akinetes
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formed as a result of uranium stress (120 h).
Figure 1
Figure 2
Figure 3