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Bacterial-facilitated uranium transport in the presence of phytate at Savannah River Site
Chemosphere 223 (2019) 351e357
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Bacterial-facilitated uranium transport in the presence of phytate at Savannah River Site Runwei Li a, Victor Ibeanusi b, Jada Hoyle-Gardner b, Christy Crandall b, Charles Jagoe b, John Seaman c, Aavudai Anandhi d, Gang Chen a, * a
Department of Civil and Environmental Engineering, FAMU-FSU College of Engineering, Tallahassee, FL, 32310, United States School of the Environment, Florida A&M University, Tallahassee, FL, 32307, United States Savannah River Ecology Laboratory, University of Georgia, Aiken, SC, 29802, United States d Department of Biological System Engineering, Florida A&M University, Tallahassee, FL, 32307, United States b c
h i g h l i g h t s UO2þ 2 transport was facilitated by indigenous bacterial strains at Savannah River Site (SRS). Facilitated UO2þ 2 transport was highly affected by its interactions with the mobile indigenous bacteria. Phytate played an important role in UO2þ 2 immobilization at Savannah River Site (SRS). Increase of pH enhanced UO2þ 2 immobilization with phytate addition.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 23 September 2018 Received in revised form 29 January 2019 Accepted 10 February 2019 Available online 13 February 2019
At the Department of Energy (DOE) managed Savannah River Site (SRS), uranium and other heavy metals continue to pose threats to the ecosystem health and processes. In the oxic soil of this site, uranium is 2þ present primarily as soluble salts of the uranyl ion (i.e., U(VI) or UO2þ 2 ). Although UO2 has a strong sorption to the soil, the mobile indigenous bacteria may facilitate its transport. On the contrary, prewith phosphate has been found to be an alternative remediation strategy. This cipitation of UO2þ 2 research investigated the effects of mobile bacteria and phytate on UO2þ 2 transport at SRS in column experiments. It was discovered that UO2þ 2 can barely be mobilized by de-ionized water but can be significantly transported with the aid of mobile indigenous bacteria. UO2þ 2 had the most facilitated transport observation when it reached equilibrium with the bacteria before the transport. When UO2þ 2 and bacterial were introduced to the soil at the same time or UO2þ 2 was pre-deposited in the soil, the facilitated transport was less pronounced. In the presence of phytate, bacterial-facilitated UO2þ 2 transport was hindered. pH was found to play the key role for UO2þ 2 immobilization in the presence of phytate. The immobilization of UO2þ 2 with the addition of phytate increased with the increase of pH within the pH range of this study because of the impact of pH on the solubility of UO2(OH)2. Phytate promoted UO-22þ PO34 complex and/or [Ca(UO2)2(PO4)2] formation, leading to enhanced UO2 immobilization in the SRS soil. Published by Elsevier Ltd.
Handling Editor: Martine Leermakers Keywords: UO2þ 2 Transport Phytate Facilitated MRS-1 Bacillus cereus
1. Introduction As many as eleven DOE facilities have handled quantities of recycled uranium (Dong et al., 2012; Li et al., 2014). At the
* Corresponding author. Department of Civil and Environmental Engineering, FAMU-FSU College of Engineering, 2525 Pottsdamer Street, Tallahassee, FL, 32310, United States. E-mail address: [email protected] (G. Chen). https://doi.org/10.1016/j.chemosphere.2019.02.064 0045-6535/Published by Elsevier Ltd.
Department of Energy (DOE) managed Savannah River Site (SRS), located along the Savannah River near Aiken, SC, radionuclides, such as uranium and heavy metals continue to pose threats to the ecosystem health and processes. In the oxic soil of this site, uranium is present primarily as soluble salts of the uranyl ion (i.e., U(VI) or UO2þ 2 ), which has a higher solubility as compared to that of anoxic soil where uranium exists as uraninite (i.e., U(IV) or UO2) (~105 M as compared to 108 M) (Buhl et al., 2005; Fortin et al., 2004). Although UO2þ 2 is considered “readily soluble” and being mobile in
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the subsurface, it is subject to adsorption to the soil surfaces which are negatively charged. UO2þ transport in the subsurface soil 2 therefore is further facilitated by mobile bacteria. The functional groups on the bacterial surfaces such as carboxyl, phosphoryl, and hydroxyl as well proteins may bind UO2þ 2 over a large pH range because of the low isoelectric point of bacterial surfaces (Kelly et al., 2001). The high surface area to volume ratio of bacteria also allows them to accumulate UO2þ 2 on their surfaces. It has been demonstrated that U(VI) can be reduced and immobilized biologically in the subsurface soil. A metabolically and phylogenetically diverse group of dissimilatory metal-reducing bacteria has been discovered to be able to reduce U(VI) to U(IV) (Truex et al., 1997). These bacteria play a key role in the subsurface of SRS since, as facultative bacteria, they are key players in the aerobic to anaerobic transition zones that are critical in defining the likely mobilization or immobilization of multivalent metals such as actinides. The U(VI)-reducing Proteobacteria primarily comprise Geobacteraceae members and mesophilic sulfate reducers within the Deltaproteobacteria, while the remainder are primarily members of the Gammaproteobacteria within the families Pseudomonadaceae and Shewanellaceae (Liu et al., 2002; Roden and Scheibe, 2005; Truex et al., 1997). From a redox perspective, microbial processes under anaerobic conditions likely reduce most higher-valent actinides. Reduction of the soluble, oxidized form of uranium, U(VI), to insoluble U(IV) such as uraninite (UO2) is an important mechanism for the immobilization of uranium at the contaminated site of SRS. Recently, precipitation of U(VI) with phosphate has been found to be an alternative remediation strategy, which has the obvious advantage that anoxic conditions are not required. To immobilize U(VI), phosphate minerals, polyphosphates or organophosphates are introduced to the subsurface to form phosphate-uranium complexation at the contaminated sites (Huynh et al., 2016; Vazquez et al., 2007). Phosphate promotes U(VI) immobilization through the formation of U(VI)ephosphate ternary surface complexes, and in the presence of calcium, the precipitation of autunite (i.e., [Ca(UO2)2(PO4)2]). Especially, autunite has extremely positive thermodynamic stability constants, producing mineral phases that are stable under common groundwater conditions. In prior research, solid-phase phosphate in the form of hydroxyapatite was used to achieve phosphate-mediated U(VI) immobilization remediation strategies (Arey et al., 1999). Organophosphates were also used to form the nonreductive precipitation of UO2þ 2 as a stable uranyl-phosphate-mineral. Presently, phytate, also known as inositol hexaphosphate (IP6), is drawing attention in remediating uranium-contaminated sites. IP6 is a major fiber-associated component with a principal role of phosphorus storage. However, phosphorus in the phytate form is not available to precipitate uranium and phytate needs to be mineralized to release phosphate in order to immobilize uranium at the contaminated sites. Since phytate is more recalcitrant to degradation than other organophosphate molecules, phytate has the advantage to be use in in-situ uranium remediation by migrating further in the contaminated subsurface soil. Indigenous bacterial strains in the uranium contaminated soil have been discovered to able to develop the capacity to mineralize phytate and release phosphate, particularly under metal-rich, phosphate-poor subsurface conditions (Horii et al., 2013; Lim et al., 2007). Prior pure-culture studies have also demonstrated the nonspecific phosphatase enzyme activities in microorganisms and microbial communities isolated from the contaminated subsurface (Suleimanova et al., 2015). These enzymes were expressed constitutively in microbes isolated from the subsurface and were responsible for the release of phosphate from organophosphates. In the contaminated subsurface, aerobic and
anaerobic (i.e., nitrate reducing) conditions co-existed and these enzymes were able to be released under both conditions under the pH range of 5.5e7. Although the presence of phytate is supposed to immobilize uranium in the contaminated soil, some research had opposite observations. For instance, one study discovered that phytate amendments increased the solubility of contaminant metals under certain conditions (Seaman et al., 2003). The objective of this research was to evaluate the effects of mobile bacteria and phytate on UO2þ 2 transport at SRS, which was conducted in column experiments. The experiments were designed to examine the impact of interactions of UO2þ 2 with the mobile indigenous bacteria in the presence of phytate on its transport. The role that pH played in above processes was also investigated. 2. Materials and methods The contaminated soil samples were collected 3 feet below the surface at the Savannah River Site (SRS) near Aiken, SC and were used for the cultivation of the uranium- and phytate-tolerate bacteria. Based on the sieving analysis, around 90.1% of the particles were found to be smaller than 0.425 mm, i.e., passing through the 40 sieve. Around 0.67% of the particles were found to be smaller than 0.075 mm, i.e., passing through the 200 sieve. In addition to soil particle size distribution, the soil hydraulic properties were characterized using the pressure-plate methods with saturated hydraulic conductivity were found to be in the range of 1.45e48.7 cm/h. The organic carbon was found to be ranged from 1.47 to 3.25% by the Mebius method (Yeomans and Bremner, 1998). The soil samples were sealed in the sealed container to prevent the loss of moisture and to minimize their exposure to the air. They were transported back to the laboratory by maintaining at 4 C immediately for continuous cultivation and enrichment. Specifically, 10 mg soil was transferred into a 250 ml Erlenmeyer flask containing 100 ml sterilized culture medium amended with 5 mg/l U(VI) and 10 mM phytate. The medium was modified from the Postgate C medium and consisted: KH2PO4, 0.5 g/l; NH4Cl, 1.0 g/l; yeast, 1.0 g/l; CaCl2$6H2O, 0.1 g/l; MgSO4$7H2O, 2.0 g/l; Na2SO4, 4.5 g/l; FeSO4$7H2O, 0.002 g/l; and 70% sodium lactate, 0.5 g/l. U(VI) was added from a stock solution of 0.01 M UO2þ 2 (in HCl). The medium pH was adjusted to 7.0 with 0.1 M NaOH and autoclaved. The inoculated Erlenmeyer flask was put into a rotary-shaker (150 rpm at room temperature) for at least 1 week until the formation of black precipitate at the bottom and on the wall of the Erlenmeyer flask could be observed. Then 10 ml enriched culture was transferred into 100 ml fresh culture medium amended with 5 mg/l U(VI) and 10 mM phytate for the second phase culture enrichment. After the fourth phase enrichment was completed, bacterial cells were harvested by centrifugation (6000 g, 15 min) and washed twice with NaHCO3 buffer (0.05 M). The concentrated bacterial cells were re-suspended in an Erlenmeyer flask containing NaHCO3 buffer (0.05 M) to give a final concentration of approximately 5 108 cells/ml. The bacterial species were identified based on polymerase chain reaction (PCR) analysis (Ibeanusi et al., 2018). Upon verification of the PCR reaction by viewing the gel bands, the PCR samples were purified using a QIAGEN QIAquick-spin PCR purification kit. After the purification, the samples were amplified, and the resulted sequences were compared with the database of the National Center for Biotechnology Information (NCBI) based on the strands that have been previously identified. The top strains whose DNA codes matched the codes of the samples with the highest certainty were identified as MRS-1 Bacillus cereus, which was deposited in ATCC (55,673).
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2.1. Column experiments Column experiments were conducted to study the MRS-1 B. cereus-facilitated UO2þ 2 transport and impact of phytate on the immobilization of UO2þ at SRS. The column experiments were 2 conducted under saturated conditions with the column (2.5 cm ID 10 cm Length) vertically oriented. The bottom of the column was sealed with a custom fit to permit the flow of water and retain the soil. To pack the column with soil samples collected from SRS, the column was first filled with deionized water (DI) water to a height of 2e3 cm and the soil samples were packed in the column through CO2 solvation to eliminate air pockets. During packing, the height of 2e3 cm of DI water was maintained. Before the introduction of phytate, UO2þ 2 transport experiments were conducted in three different ways of UO2þ 2 introduction to 2þ examine the impact of UO2þ 2 -bacteria interactions on UO2 transport. Scenario 1: MRS-1 B. cereus and UO2þ were mixed to reach 2 equilibrium and then the mixture was introduced to the column by a peristaltic pump (Masterflex, Cole-Parmer, Vernon Hills, IL). The column was then flushed with DI water and MRS-1 B. cereus suspension. Scenario 2: MRS-1 B. cereus and UO2þ 2 were introduced to the column spontaneously from separate reservoirs. The column was then flushed with DI water and MRS-1 B. cereus suspension. Scenario 3: UO2þ 2 was introduced and deposited in the column first. The column was then flushed with DI water and MRS-1 B. cereus suspension. For all above experiments, UO2þ was applied at a 2 concentration of 10 mg/l and MRS-1 B. cereus at a concentration of 5 108 cells/ml. Elution was collected by a fraction collector and UO2þ 2 was measured by the Microwave Plasma-Atomic Emission Spectrophotometer and MRS-1 B. cereus was quantified by a UVevis spectrometry. 2þ To study the impact of phytate on UO2þ 2 immobilization, UO2 (10 mg/l), MRS-1 B. cereus (5 108 cells/ml) and phytate (10 mM) were introduced in the column. The column was left untouched for 10 days to allow MRS-1 B. cereus to release phosphate from organophosphates and interact with UO2þ 2 . 10 days later, the column was flushed with DI water first. When background UO2þ was 2 efluted from the column, the column was then flushed with MRS-1 B cereus suspension. DI water flushing was to investigate the possible mobilization of UO2þ by groundwater flow alone. The 2 MRS-1 B cereus flushing was to further investigate the mobilization of UO2þ 2 with the aid of mobile MRS-1 B cereus. Parallel column experiments in the absence of phytate were conducted as the controls. Above experiments were conducted at pH 4.2, 6.8 and 8.2, respectively with pH adjusted by 0.1 M HCl or 0.1 M NaOH. For each series of the column experiments, a fresh column was packed and prior to starting each experiment. The column had a volume of 31.4 ml with a pore volume of 17.3 ml. 173 ml (i.e., 10 pore volumes) of DI water was eluted through the column to stabilize the column. A conservative pulse tracer (chloride) breakthrough curve was generated separately to characterize the soil. UO2þ 2 transport in the SRS soil was described by the equilibriumkinetic two-site model (Toride et al., 1995):
bR
vC1 1 v2 C1 vC1 ¼ $ 2 uðC1 C2 Þ m1 C1 P vZ vT vZ
ð1 bÞR
vC2 ¼ uðC1 C2 Þ m2 C2 vT
where C1 and C2 are the dimensionless UO2þ 2 concentration in the solution and on the soil surface, respectively; b is the partition coefficient; R is the retardation factor (R ¼ 1 þ rbqKd , where rb is the soil bulk density, Kd is the partition coefficient of UO2þ 2 between the
353
solution and the soil, and q is the soil porosity); T is the dimensionless time; P is the Peclet number (P ¼ vL D , where v is the interstitial pore-water velocity, L is the length of the column, and D is the dispersion coefficient); Z is the dimensionless axial coordinate; u is the dimensionless mass transfer coefficient; and m1 and m2 are the dimensionless deposition coefficient in the solution and on the soil surface, respectively. For this research, it was assumed that the dominating UO2þ 2 deposition occurred from the solution. Subsequently, m2 was set to 0. For UO2þ transport, a pulse-type 2 boundary condition was used for the upper boundary and a zero gradient was assumed for the lower boundary (Reed, 1965):
qD
vC1 þ qC1 vZ
¼
Z¼0
qC0 0 < t t0 0 t > t0
vC1 ¼0 vZ Z¼L where q is the specific discharge (Darcian fluid flux), t0 is the duration of injection, and C0 is the initial UO2þ concentration. 2 Transport parameters in above equations were obtained by fitting the experimentally obtained UO2þ breakthrough data using an 2 implicit, finite-difference scheme of CXTFIT (Toride et al., 1995). All the parameters were optimized by minimizing the sum of squared differences between observed and fitted concentrations using the nonlinear least-square method (Toride et al., 1995). The simulated transport parameters are summarized in Table 1. It should be noted that above parameters were in the unit-less format. For deposition coefficient, m1, the corresponding values reported later were 1.5 103 min1, 9.2 103 min1 and 1.2 102 min1 for Scenario 1, 2 and 3, respectively. Effect of bacteria cannot be directly reflected by these models. On the other hand, the facilitated transport was reflected indirectly by the comparison of the simulated transport results with or without the aid of the facilitating bacteria.
3. Results and discussion Tracer (Cl) transport was studied before MRS-1 B. cereusfacilitated UO2þ 2 transport experiments in the SRS soil. Nearly all the input tracer was eluted from the column (Fig. 1). The tracer breakthrough curve was characterized by a breakthrough front and an elution tail and was simulated against transport equations. During the model simulation, the retardation factor was set to 1.0, i.e., Kd ¼ 0. This was based on the consideration that the tracer should not be retarded in the soil as the tracer was assumed not to adsorb in the soil. In addition, the deposition coefficient was set to zero, i.e., no retention of the tracer in the soil. This was true since nearly all the inputted tracer was eluted from the column at the end of the transport experiments. After the simulation, D was determined to be 1.30 cm2/min and then used for the simulations of UO2þ 2 transport. Breakthrough curves of MRS-1 B. cereus-facilitated UO2þ 2 transport in the SRS soil were characterized by a self-sharpening front, which became broader and diffuser at the elution limb (Fig. 1). The
Table 1 UO2þ 2 transport parameters simulated by the equilibrium-kinetic two-site model.
Chloride Scenario 1 Scenario 2 Scenario 3
b
u
m1
m2
0.990 ± 0.11 0.992 ± 0.08 0.998 ± 0.06 0.987 ± 0.12
0.02 ± 0.04 0.08 ± 0.05 1.43 ± 0.45 1.75 ± 0.61
Set to 0.00 5.2 105 ± 1.0 105 3.1 104 ± 1.5 105 4.2 104 ± 2.7 105
Set Set Set Set
to to to to
0.00 0.00 0.00 0.00
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2þ Fig. 1. UO2þ 2 breakthrough curves for different UO2 introduction scenarios.
long-lasting tails of the breakthrough curves indicated kineticcontrolled UO2þ 2 retention in the column. Under saturated conditions, UO2þ transport is controlled by kinetic adsorption and 2 equilibrium adsorption processes. The irreversible adsorption or kinetic adsorption in the SRS soil was evidenced by the reduced mass recovery (the integration of the breakthrough curve) during the transport and the equilibrium adsorption was demonstrated by delayed breakthrough reflected by the retardation factor. The UO2þ 2 breakthrough curves were simulated against the transport equations by CXTFIT (Toride et al., 1995). UO2þ 2 breakthrough curves were fitted well with the model and the accuracy of transport modeling was expressed by the sum of the squared differences between observed and fitted data. The mean square for error was in the range from 2.97 104 to 1.21 102. With the aid of MRS-1 B. cereus, UO2þ 2 transport was significantly facilitated by increased mass recovery of UO2þ 2 that passed through the column. Among the three scenarios, Scenario 1, i.e.,
UO2þ 2 pre-equilibriumed with MRS-1 B. cereus had the most facilitated transport observation with greatest UO2þ 2 mass recovery as compared to that of UO2þ 2 transport alone. For the other two scenarios, although increased mass recovery of UO2þ 2 was observed with the aid of the mobile MRS-1 B. cereus, the increase was much less pronounced. UO2þ 2 retention in the SRS soil was described by the deposition coefficient. With the aid of MRS-1 B. cereus, the corresponding deposition coefficient values decreased from 2 5.3 102 min1 for UO2þ min1, 2 transport alone to 1.2 10 9.2 103 min1 and 1.5 103 min1 for Scenario 3, 2 and 1, respectively. These deposition coefficient values reflected the irreversible adsorption of UO2þ 2 to the SRS soil and inversely corresponded to mass recovery of UO2þ 2 that passed through the column. The different deposition coefficient values of UO2þ 2 in the SRS soil for different facilitation scenarios indicated that the UO2þ in2 teractions with MRS-1 B. cereus affected the facilitated transport of UO2þ 2 . By comparing the breakthrough curves, it was discovered that for the three scenarios, UO2þ 2 breakthrough coincided with MRS-1 B. cereus transport (Fig. 2). This evidenced that UO2þ 2 transport was facilitated by MRS-1 B. cereus. The different bacterial transport observations among different scenarios were attributed to the formation of UO2þ 2 -bacteria complexation. For different scenarios, the different formation of UO2þ 2 -bacteria complexation impacted bacterial surface properties and the subsequently bacterial transport. The formation of different UO2þ 2 -bacteria complexation also affected UO2þ 2 transport. For Scenario 1, equilibrium was reached and the most complete UO2þ 2 -bacteria complexation was achieved, leading the most significant facilitated UO2þ 2 transport. For Scenario 2 and Scenario 3, less complete UO2þ 2 -bacteria complexation formation led to less significant facilitated UO2þ 2 transport. For Scenario 1 and Scenario 2, UO2þ 2 was carried by MRS-1 B. cereus during the transport in the soil, which also suffered from desorption and re-adsorption interactions between MRS-1 B. cereus and the soil. However, for Scenario 3, UO2þ 2 was desorbed from the SRS soil first and then transported with facilitating MRS-1 B. cereus. The ratio of UO2þ 2 /MRS-1 B. cereus eluted from the column as a function of pore volume clearly demonstrated different facilitated UO2þ 2
Fig. 2. UO2þ 2 /MRS-1 B. cereus. Ratio when eluted from the column.
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observations for different UO2þ introduction scenarios (Fig. 2). 2 Among these three scenarios, Scenario 1, i.e., UO2þ pre2 equilibriumed with MRS-1 B. cereus had the highest ratio, Scenario 3, i.e., UO2þ 2 pre-deposited in the soil column had the lowest ratio, and Scenario 2, i.e., UO2þ 2 introduced spontaneously with MRS-1 B. cereus was in between. When transported in the SRS soil, UO2þ 2 was also retarded as evidenced by the delayed breakthrough front, which was described by the retardation factor. Theoretically, retardation can be estimated from the breakthrough curve by (Kim and Kim, 2007; Li et al., 2009):
ð∞
C tdt tp C R ¼ ð0∞ 0 C 2 dt C 0 0 where t is the pore volume and tp is pulse width or pulse duration in pore volumes as defined by Dohse and Lion (1994). MRS-1 B. cereus facilitation also decreased UO2þ 2 retardation in the SRS soil. Retardation factor is the indication of the relative velocity of 2þ UO2þ 2 to that of water. UO2 transport was facilitated by the mobile bacteria, which was reflected by the retardation factor. The retardation factors obtained based on above equation were 4.04, 4.48 and 6.29 for Scenario 1, 2 and 3, respectively, which meant that water transport velocity was 4.04, 4.48 and 6.29 times that of UO2þ 2 . Above observations indicated that UO2þ 2 transport was most facilitated for Scenario 1 and least facilitated for Scenario 3. For the study of impact of phytate on UO2þ 2 immobilization, UO2þ and phytate were pre-deposited in the soil column first 2 before flushing with DI water and MRS-1 B. cereus suspension. The introduction of UO2þ 2 under different pH conditions was consistent with UO2þ 2 breakthrough curves almost overlapped with those of the control experiments (Fig. 3). Based on the mass balance, 0.038 mg of UO2þ 2 was pre-deposited in 8.09 g soil in the column. After 10 days, the column was flushed with DI water, followed by MRS-1 B. cereus suspension. During the 10 days’ period of time, MRS-1 B. cereus decomposed phytate to release phosphate. With the available phosphate and possible calcium, variable unsolvable UO-2-PO34 complexes and autunite [Ca(UO2)2(PO4)2] precipitation could be formed. The un-precipitated portion of UO2þ 2 was also adsorbed to the negative charged soil surface. Subsequently, minimal UO2þ 2 was flushed out by DI water (Fig. 3). However, when flushed by MRS-1 B. cereus suspension, the adsorbed UO2þ 2 may desorb from the soil and carried by MRS-1 B. cereus to break through the column. In addition, the UO-2-PO34 complexes and Ca(UO2)2(PO4)2 may also be possible to be carried out by mobile MRS-1 B. cereus. In another research, 10 mM organophosphate amendment was used between pH 4.5 and 7.0 and 73e95% of total uranium was precipitated after 120 h of incubation. This research demonstrated that indigenous bacterial within the uraniumcontaminated environment can hydrolyze organophosphate, especially under low pH conditions to immobilize UO2þ 2 (Beazley et al., 2007). The results were consistent with our observation that minimal UO2þ 2 was flushed out by DI water. Research using glycerol phosphate had similar observations that indigenous soil bacteria promoted the precipitation of UO2þ 2 at pH 5.5 and 7.0 (Beazley et al., 2011; Martinez et al., 2014; Salome et al., 2013). Phytate played an important role for the MRS-1 B. cereus-facil2þ itated UO2þ 2 transport as evidenced by the fact that much less UO2 was able to be flushed out by MRS-1 B. cereus with the addition of phytate as compared to that of the control. Facilitated UO2þ 2 transport by MRS-1 B. cereus with phytate addition was also a function of solution pH. With the increase of pH, less UO2þ 2 was flushed out by MRS-1 B. cereus (Fig. 3). To understand the impact of
Fig. 3. UO2þ 2 breakthrough curves by DI water and MRS-1 B. cereus. Suspension after 10 days of incubation with phytate.
pH on MRS-1 B. cereus-facilitated UO2þ transport with phytate 2 addition, the graph of uranium solubility versus pH was con4þ structed based on the introduced UO2þ 2 , released PO4 and pore water calcium concentration. Uranium solubility decreased with the increase of pH until pH 8, and then increased with the continuous increase of pH (Fig. 4). Because of the high solubility at pH 4, UO2þ 2 was easier to be transported with the facilitation of MRS-1 B. cereus. On the contrary, due to the low solubility at pH 8.2, less UO2þ was flushed out resulting from UO2þ precipitation. 2 2 Assuming a first order release from the SRS soil (Mishra et al., 2012; Rahmani-Sani et al., 2015), the release coefficient linearly decreased with the increase of pH (Fig. 5).
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Fig. 4. UO2þ 2 Speciation as a Function of pH.
Fig. 5. UO2þ 2 Release Coefficient as a Function of pH.
UO2þ 2 adsorption in the SRS soil was estimated by integrating the diffuse front of the breakthrough curves (Burgisser et al., 1993) (Fig. 6). Based on this method, the concentration of equilibrium adsorbed UO2þ 2 in the SRS soil, S, can be obtained by integrating the experimental record of the retention time t(c) if the dispersion term can be neglected (D ¼ 0): Fig. 6. UO2þ 2 sorption isotherms in SRS soil.
S¼
q rb ð1 qÞ
ðC 0
ðC tðC 0 Þ 0 1 dC tðC 0 ÞKc dC 0 t0 0
where t0 ¼ Lv and is the average UO2þ 2 travel time in the column (min). The insignificant role of hydrodynamic dispersion on UO2þ 2 transport has been proven to be valid (Unice and Logan, 2000). In their research, they demonstrated that hydrodynamic dispersion can be neglected for Pe > 100. For this research, all the column experiments were performed with Pe > 100, thus above equation can be used to determine the UO2þ 2 adsorption isotherms. The adsorption of UO2þ 2 in the SRS soil increased with the increase of the equilibrium concentration. UO2þ had greater 2 adsorption in the SRS soil with the phytate addition than those of without, implying UO-2-PO34 complexes and/or [Ca(UO2)2(PO4)2] precipitation occurred, which increased its sorption in the SRS soil.
UO2þ 2 adsorption in the SRS soil was described by Langmuir isotherms (Adachi et al., 1997; Hajratwala, 1982):
S¼
qmax KCe 1 þ KCe
where S is the adsorbed UO2þ 2 in the SRS soil; qmax is the maximum mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface, Ce is the UO2þ 2 equilibrium concentration, and K is the Langmuir adsorption constant, which increases with the increase of the binding energy of adsorption. UO2þ 2 had qmax values of 0.012, 0.015 and 0.020 mM/g for pH 4.2, 6.8 and 8.2 without the addition of phytate. With the addition phytate, these values increased to 0.020, 0.028 and 0.038 mM/g, respectively. qmax is the maximum
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mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface. Above observations indicated that the increase of pH increased the mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface. In the presence of phytate, the mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface further increased. The underlying principle behind these observations was the formation of bonding between UO2þ 2 and adsorption receptor sites on SRS soil surfaces. Increase of pH and the presence of phytate increased the bonding energy between UO2þ 2 and the adsorption receptor sites on SRS soil surfaces. 4. Conclusions Under oxic conditions at the uranium-contaminated sites, uranium is primarily in dissolved form of UO2þ 2 , which can get equilibrium with indigenous MRS-1 B. cereus. The subsequent MRS-1 B. cereus transport can facilitate UO2þ 2 transport. This facilitated transport may also occur before equilibrium is reached. In addition, UO2þ 2 can be adsorbed to the soil before MRS-1 B. cereus transport occurs. These three possible scenarios of MRS-1 B. cereus-facilitated UO2þ transport were investigated in this research. In practice, 2 addition of phytate has been practiced as a means of UO2þ 2 immobilization. The presence of phytate on MRS-1 B. cereus-facilitated UO2þ 2 transport was also explored. In this research, transport of UO2þ 2 in the SRS soil was found to be significantly facilitated by MRS-1 B. cereus. The facilitation results were controlled by UO2þ 2 interactions with MRS-1 B. cereus. Pre-equilibrium with MRS-1 B. cereus before the introduction had the most facilitated transport observations. Further analysis showed that MRS-1 B. cereusfacilitated transport was due to the adsorption of UO2þ 2 on the MRS1 B. cereus surface based on the quantification of the ratio of UO2þ 2 / MRS-1 B. cereus eluted from the column. Phytate played an important role for the MRS-1 B. cereus-facilitated UO2þ 2 transport, which was found to be a function of solution pH. The immobilization of UO2þ 2 with the addition of phytate increased with the increase of pH within the pH range of this study because uranium solubility decreased with the increase of pH until pH 8. Acknowledgements The work was supported by Department of Energy Minority Serving Institution Partnership Program (MSIPP) managed by the Savannah River National Laboratory under SRNS contract DE-AC0908SR22470. Appendix A. Supplementary data Supplementary data related to this article can be found at https://doi.org/10.1016/j.chemosphere.2019.02.064. References Adachi, S., Panintrarux, C., Matsuno, R., 1997. Methods for estimating the parameters of nonlinear adsorption isotherms of Langmuir and Freundlich types from a response curve of pulse input of an adsorbate. Biosci. Biotechnol. Biochem. 61, 1626e1633. Arey, J.S., Seaman, J.C., Bertsch, P.M., 1999. Immobilization of uranium in contaminated sediments by hydroxyapatite addition. Environ. Sci. Technol. 33, 337e342. Beazley, M.J., Martinez, R.J., Sobecky, P.A., Webb, S.M., Taillefert, M., 2007. Uranium biomineralization as a result of bacterial phosphatase activity: insights from bacterial isolates from a contaminated subsurface. Environ. Sci. Technol. 41, 5701e5707. Beazley, M.J., Martinez, R.J., Webb, S.M., Sobecky, P.A., Taillefert, M., 2011. The effect of pH and natural microbial phosphatase activity on the speciation of uranium in subsurface soils. Geochem. Cosmochim. Acta 75, 5648e5663. Buhl, M., Diss, R., Wipff, G., 2005. Coordination environment of aqueous uranyl (VI)
357
ion. J. Am. Chem. Soc. 127, 13506e13507. Burgisser, C.S., Cernik, M., Borkovec, M., Sticher, H., 1993. Determination of nonlinear adsorption-isotherms from column experiments: an alternative to batch studies. Environ. Sci. Technol. 27, 943e948. Dohse, D.M., Lion, L.W., 1994. Effect of microbial polymers on the sorption and transport of phenanthrene in a low-carbon sand. Environ. Sci. Technol. 28, 541e548. Dong, W., Tokunaga, T.K., Davis, J.A., Wan, J., 2012. Uranium (VI) adsorption and surface complexation modeling onto background sediments from the F-area Savannah River Site. Environ. Sci. Technol. 46, 1565e1571. Fortin, C., Dutel, L., Garnier-Laplace, J., 2004. Uranium complexation and uptake by a green alga in relation to chemical speciation: the importance of the free uranyl ion. Environ. Toxicol. Chem. 23, 974e981. Hajratwala, B.R., 1982. Potential errors in determining Freundlich and Langmuir constants from adsorption isotherms. J. Pharmacol. Sci. 71, 125e126. Horii, S., Matsuno, T., Tagomori, J., Mukai, M., Adhikari, D., Kubo, M., 2013. Isolation and identification of phytate-degrading bacteria and their contribution to phytate mineralization in soil. J. Gen. Appl. Microbiol. 59, 353e360. Huynh, T.S., Vidaud, C., Hagege, A., 2016. Investigation of uranium interactions with calcium phosphate-binding proteins using ICP/MS and CE-ICP/MS. Metall 8, 1185e1192. Ibeanusi, V., Pathak, A., Chauhan, A., Hoyle-Gardner, J., Cooper, T., Turker, L., Howard, H., Obinegbo, O., Chen, G., Seaman, J., 2018. Genome-centric evaluation of Bacillus sp. strain: ATCC55673 and response to uranium biomineralization. Significance Bioeng Biosci 2, 1e8. Kelly, S.D., Boyanov, M.I., Bunker, B.A., Fein, J.B., Fowle, D.A., Yee, N., Kemner, K.M., 2001. XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J. Synchrotron Radiat. 8, 946e948. Kim, M., Kim, S.B., 2007. Modeling contaminant transport in a three-phase groundwater system with the Freundlich-type retardation factor. Environ. Technol. 28, 205e216. Li, D., Seaman, J.C., Chang, H.S., Jaffe, P.R., Koster van Groos, P., Jiang, D.T., Chen, N., Lin, J., Arthur, Z., Pan, Y., Scheckel, K.G., Newville, M., Lanzirotti, A., Kaplan, D.I., 2014. Retention and chemical speciation of uranium in an oxidized wetland sediment from the Savannah River Site. J. Environ. Radioact. 131, 40e46. Li, M.H., Wang, T.H., Teng, S.P., 2009. Experimental and numerical investigations of effect of column length on retardation factor determination: a case study of cesium transport in crushed granite. J. Hazard Mater. 162, 530e535. Lim, B.L., Yeung, P., Cheng, C., Hill, J.E., 2007. Distribution and diversity of phytatemineralizing bacteria. ISME J. 1, 321e330. Liu, C., Gorby, Y.A., Zachara, J.M., Fredrickson, J.K., Brown, C.F., 2002. Reduction kinetics of Fe(III), Co(III), U(VI), Cr(VI), and Tc(VII) in cultures of dissimilatory metal-reducing bacteria. Biotechnol. Bioeng. 80, 637e649. Martinez, R.J., Wu, C.H., Beazley, M.J., Andersen, G.L., Conrad, M.E., Hazen, T.C., Taillefert, M., Sobecky, P.A., 2014. Microbial community responses to organophosphate substrate additions in contaminated subsurface sediments. PLoS One 9, e100383. Mishra, S., Maity, S., Bhalke, S., Pandit, G.G., Puranik, V.D., Kushwaha, H.S., 2012. Thermodynamic and kinetic investigations of uranium adsorption on soil. J. Radioanal. Nucl. Chem. 294, 97e102. Rahmani-Sani, A., Hosseini-Bandegharaei, A., Hosseini, S.H., Kharghani, K., Zarei, H., Rastegar, A., 2015. Kinetic, equilibrium and thermodynamic studies on sorption of uranium and thorium from aqueous solutions by a selective impregnated resin containing carminic acid. J. Hazard Mater. 286, 152e163. Reed, K.W., 1965. A time dependent boundary condition for neutron transport equation. Am. Math. Mon. 72, 454. Roden, E.E., Scheibe, T.D., 2005. Conceptual and numerical model of uranium(VI) reductive immobilization in fractured subsurface sediments. Chemosphere 59, 617e628. Salome, K.R., Green, S.J., Beazley, M.J., Webb, S.M., Kostka, J.E., Taillefert, M., 2013. The role of anaerobic respiration in the immobilization of uranium through biomineralization of phosphate minerals. Geochem. Cosmochim. Acta 106, 344e363. Seaman, J.C., Hutchison, J.M., Jackson, B.P., Vulava, V.M., 2003. In situ treatment of metals in contaminated soils with phytate. J. Environ. Qual. 32, 153e161. Suleimanova, A.D., Beinhauer, A., Valeeva, L.R., Chastukhina, I.B., Balaban, N.P., Shakirov, E.V., Greiner, R., Sharipova, M.R., 2015. Novel glucose-1-phosphatase with high phytase activity and unusual metal ion activation from soil bacterium Pantoea sp. Strain 3.5.1. Appl. Environ. Microbiol. 81, 6790e6799. Toride, N.L., Leij, F.J., van Genuchten, M. Th., 1995. The CXTFIT Code for Estimating Transport Parameters from Laboratory or Field Experiments, Version 2.1. U.S. Salinity Laboratory, Riverside, CA. Truex, M.J., Peyton, B.M., Valentine, N.B., Gorby, Y.A., 1997. Kinetics of U (VI) reduction by a dissimilatory Fe (III)-reducing bacterium under non-growth conditions. Biotechnol. Bioeng. 55, 490e496. Unice, K.M., Logan, B.E., 2000. Insignificant Role of hydrodynamic dispersion on bacterial transport. J. Environ. Eng. 126, 491e500. Vazquez, G.J., Dodge, C.J., Francis, A.J., 2007. Interactions of uranium with polyphosphate. Chemosphere 70, 263e269. Yeomans, J.C., Bremner, J.M., 1998. A rapid and precise method for routine determination of organic carbon in soil. Commun. Soil Sci. Plant Anal. 19, 1467e1476.
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Bacterial-facilitated uranium transport in the presence of phytate at Savannah River Site Runwei Li a, Victor Ibeanusi b, Jada Hoyle-Gardner b, Christy Crandall b, Charles Jagoe b, John Seaman c, Aavudai Anandhi d, Gang Chen a, * a
Department of Civil and Environmental Engineering, FAMU-FSU College of Engineering, Tallahassee, FL, 32310, United States School of the Environment, Florida A&M University, Tallahassee, FL, 32307, United States Savannah River Ecology Laboratory, University of Georgia, Aiken, SC, 29802, United States d Department of Biological System Engineering, Florida A&M University, Tallahassee, FL, 32307, United States b c
h i g h l i g h t s UO2þ 2 transport was facilitated by indigenous bacterial strains at Savannah River Site (SRS). Facilitated UO2þ 2 transport was highly affected by its interactions with the mobile indigenous bacteria. Phytate played an important role in UO2þ 2 immobilization at Savannah River Site (SRS). Increase of pH enhanced UO2þ 2 immobilization with phytate addition.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 23 September 2018 Received in revised form 29 January 2019 Accepted 10 February 2019 Available online 13 February 2019
At the Department of Energy (DOE) managed Savannah River Site (SRS), uranium and other heavy metals continue to pose threats to the ecosystem health and processes. In the oxic soil of this site, uranium is 2þ present primarily as soluble salts of the uranyl ion (i.e., U(VI) or UO2þ 2 ). Although UO2 has a strong sorption to the soil, the mobile indigenous bacteria may facilitate its transport. On the contrary, prewith phosphate has been found to be an alternative remediation strategy. This cipitation of UO2þ 2 research investigated the effects of mobile bacteria and phytate on UO2þ 2 transport at SRS in column experiments. It was discovered that UO2þ 2 can barely be mobilized by de-ionized water but can be significantly transported with the aid of mobile indigenous bacteria. UO2þ 2 had the most facilitated transport observation when it reached equilibrium with the bacteria before the transport. When UO2þ 2 and bacterial were introduced to the soil at the same time or UO2þ 2 was pre-deposited in the soil, the facilitated transport was less pronounced. In the presence of phytate, bacterial-facilitated UO2þ 2 transport was hindered. pH was found to play the key role for UO2þ 2 immobilization in the presence of phytate. The immobilization of UO2þ 2 with the addition of phytate increased with the increase of pH within the pH range of this study because of the impact of pH on the solubility of UO2(OH)2. Phytate promoted UO-22þ PO34 complex and/or [Ca(UO2)2(PO4)2] formation, leading to enhanced UO2 immobilization in the SRS soil. Published by Elsevier Ltd.
Handling Editor: Martine Leermakers Keywords: UO2þ 2 Transport Phytate Facilitated MRS-1 Bacillus cereus
1. Introduction As many as eleven DOE facilities have handled quantities of recycled uranium (Dong et al., 2012; Li et al., 2014). At the
* Corresponding author. Department of Civil and Environmental Engineering, FAMU-FSU College of Engineering, 2525 Pottsdamer Street, Tallahassee, FL, 32310, United States. E-mail address: [email protected] (G. Chen). https://doi.org/10.1016/j.chemosphere.2019.02.064 0045-6535/Published by Elsevier Ltd.
Department of Energy (DOE) managed Savannah River Site (SRS), located along the Savannah River near Aiken, SC, radionuclides, such as uranium and heavy metals continue to pose threats to the ecosystem health and processes. In the oxic soil of this site, uranium is present primarily as soluble salts of the uranyl ion (i.e., U(VI) or UO2þ 2 ), which has a higher solubility as compared to that of anoxic soil where uranium exists as uraninite (i.e., U(IV) or UO2) (~105 M as compared to 108 M) (Buhl et al., 2005; Fortin et al., 2004). Although UO2þ 2 is considered “readily soluble” and being mobile in
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the subsurface, it is subject to adsorption to the soil surfaces which are negatively charged. UO2þ transport in the subsurface soil 2 therefore is further facilitated by mobile bacteria. The functional groups on the bacterial surfaces such as carboxyl, phosphoryl, and hydroxyl as well proteins may bind UO2þ 2 over a large pH range because of the low isoelectric point of bacterial surfaces (Kelly et al., 2001). The high surface area to volume ratio of bacteria also allows them to accumulate UO2þ 2 on their surfaces. It has been demonstrated that U(VI) can be reduced and immobilized biologically in the subsurface soil. A metabolically and phylogenetically diverse group of dissimilatory metal-reducing bacteria has been discovered to be able to reduce U(VI) to U(IV) (Truex et al., 1997). These bacteria play a key role in the subsurface of SRS since, as facultative bacteria, they are key players in the aerobic to anaerobic transition zones that are critical in defining the likely mobilization or immobilization of multivalent metals such as actinides. The U(VI)-reducing Proteobacteria primarily comprise Geobacteraceae members and mesophilic sulfate reducers within the Deltaproteobacteria, while the remainder are primarily members of the Gammaproteobacteria within the families Pseudomonadaceae and Shewanellaceae (Liu et al., 2002; Roden and Scheibe, 2005; Truex et al., 1997). From a redox perspective, microbial processes under anaerobic conditions likely reduce most higher-valent actinides. Reduction of the soluble, oxidized form of uranium, U(VI), to insoluble U(IV) such as uraninite (UO2) is an important mechanism for the immobilization of uranium at the contaminated site of SRS. Recently, precipitation of U(VI) with phosphate has been found to be an alternative remediation strategy, which has the obvious advantage that anoxic conditions are not required. To immobilize U(VI), phosphate minerals, polyphosphates or organophosphates are introduced to the subsurface to form phosphate-uranium complexation at the contaminated sites (Huynh et al., 2016; Vazquez et al., 2007). Phosphate promotes U(VI) immobilization through the formation of U(VI)ephosphate ternary surface complexes, and in the presence of calcium, the precipitation of autunite (i.e., [Ca(UO2)2(PO4)2]). Especially, autunite has extremely positive thermodynamic stability constants, producing mineral phases that are stable under common groundwater conditions. In prior research, solid-phase phosphate in the form of hydroxyapatite was used to achieve phosphate-mediated U(VI) immobilization remediation strategies (Arey et al., 1999). Organophosphates were also used to form the nonreductive precipitation of UO2þ 2 as a stable uranyl-phosphate-mineral. Presently, phytate, also known as inositol hexaphosphate (IP6), is drawing attention in remediating uranium-contaminated sites. IP6 is a major fiber-associated component with a principal role of phosphorus storage. However, phosphorus in the phytate form is not available to precipitate uranium and phytate needs to be mineralized to release phosphate in order to immobilize uranium at the contaminated sites. Since phytate is more recalcitrant to degradation than other organophosphate molecules, phytate has the advantage to be use in in-situ uranium remediation by migrating further in the contaminated subsurface soil. Indigenous bacterial strains in the uranium contaminated soil have been discovered to able to develop the capacity to mineralize phytate and release phosphate, particularly under metal-rich, phosphate-poor subsurface conditions (Horii et al., 2013; Lim et al., 2007). Prior pure-culture studies have also demonstrated the nonspecific phosphatase enzyme activities in microorganisms and microbial communities isolated from the contaminated subsurface (Suleimanova et al., 2015). These enzymes were expressed constitutively in microbes isolated from the subsurface and were responsible for the release of phosphate from organophosphates. In the contaminated subsurface, aerobic and
anaerobic (i.e., nitrate reducing) conditions co-existed and these enzymes were able to be released under both conditions under the pH range of 5.5e7. Although the presence of phytate is supposed to immobilize uranium in the contaminated soil, some research had opposite observations. For instance, one study discovered that phytate amendments increased the solubility of contaminant metals under certain conditions (Seaman et al., 2003). The objective of this research was to evaluate the effects of mobile bacteria and phytate on UO2þ 2 transport at SRS, which was conducted in column experiments. The experiments were designed to examine the impact of interactions of UO2þ 2 with the mobile indigenous bacteria in the presence of phytate on its transport. The role that pH played in above processes was also investigated. 2. Materials and methods The contaminated soil samples were collected 3 feet below the surface at the Savannah River Site (SRS) near Aiken, SC and were used for the cultivation of the uranium- and phytate-tolerate bacteria. Based on the sieving analysis, around 90.1% of the particles were found to be smaller than 0.425 mm, i.e., passing through the 40 sieve. Around 0.67% of the particles were found to be smaller than 0.075 mm, i.e., passing through the 200 sieve. In addition to soil particle size distribution, the soil hydraulic properties were characterized using the pressure-plate methods with saturated hydraulic conductivity were found to be in the range of 1.45e48.7 cm/h. The organic carbon was found to be ranged from 1.47 to 3.25% by the Mebius method (Yeomans and Bremner, 1998). The soil samples were sealed in the sealed container to prevent the loss of moisture and to minimize their exposure to the air. They were transported back to the laboratory by maintaining at 4 C immediately for continuous cultivation and enrichment. Specifically, 10 mg soil was transferred into a 250 ml Erlenmeyer flask containing 100 ml sterilized culture medium amended with 5 mg/l U(VI) and 10 mM phytate. The medium was modified from the Postgate C medium and consisted: KH2PO4, 0.5 g/l; NH4Cl, 1.0 g/l; yeast, 1.0 g/l; CaCl2$6H2O, 0.1 g/l; MgSO4$7H2O, 2.0 g/l; Na2SO4, 4.5 g/l; FeSO4$7H2O, 0.002 g/l; and 70% sodium lactate, 0.5 g/l. U(VI) was added from a stock solution of 0.01 M UO2þ 2 (in HCl). The medium pH was adjusted to 7.0 with 0.1 M NaOH and autoclaved. The inoculated Erlenmeyer flask was put into a rotary-shaker (150 rpm at room temperature) for at least 1 week until the formation of black precipitate at the bottom and on the wall of the Erlenmeyer flask could be observed. Then 10 ml enriched culture was transferred into 100 ml fresh culture medium amended with 5 mg/l U(VI) and 10 mM phytate for the second phase culture enrichment. After the fourth phase enrichment was completed, bacterial cells were harvested by centrifugation (6000 g, 15 min) and washed twice with NaHCO3 buffer (0.05 M). The concentrated bacterial cells were re-suspended in an Erlenmeyer flask containing NaHCO3 buffer (0.05 M) to give a final concentration of approximately 5 108 cells/ml. The bacterial species were identified based on polymerase chain reaction (PCR) analysis (Ibeanusi et al., 2018). Upon verification of the PCR reaction by viewing the gel bands, the PCR samples were purified using a QIAGEN QIAquick-spin PCR purification kit. After the purification, the samples were amplified, and the resulted sequences were compared with the database of the National Center for Biotechnology Information (NCBI) based on the strands that have been previously identified. The top strains whose DNA codes matched the codes of the samples with the highest certainty were identified as MRS-1 Bacillus cereus, which was deposited in ATCC (55,673).
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2.1. Column experiments Column experiments were conducted to study the MRS-1 B. cereus-facilitated UO2þ 2 transport and impact of phytate on the immobilization of UO2þ at SRS. The column experiments were 2 conducted under saturated conditions with the column (2.5 cm ID 10 cm Length) vertically oriented. The bottom of the column was sealed with a custom fit to permit the flow of water and retain the soil. To pack the column with soil samples collected from SRS, the column was first filled with deionized water (DI) water to a height of 2e3 cm and the soil samples were packed in the column through CO2 solvation to eliminate air pockets. During packing, the height of 2e3 cm of DI water was maintained. Before the introduction of phytate, UO2þ 2 transport experiments were conducted in three different ways of UO2þ 2 introduction to 2þ examine the impact of UO2þ 2 -bacteria interactions on UO2 transport. Scenario 1: MRS-1 B. cereus and UO2þ were mixed to reach 2 equilibrium and then the mixture was introduced to the column by a peristaltic pump (Masterflex, Cole-Parmer, Vernon Hills, IL). The column was then flushed with DI water and MRS-1 B. cereus suspension. Scenario 2: MRS-1 B. cereus and UO2þ 2 were introduced to the column spontaneously from separate reservoirs. The column was then flushed with DI water and MRS-1 B. cereus suspension. Scenario 3: UO2þ 2 was introduced and deposited in the column first. The column was then flushed with DI water and MRS-1 B. cereus suspension. For all above experiments, UO2þ was applied at a 2 concentration of 10 mg/l and MRS-1 B. cereus at a concentration of 5 108 cells/ml. Elution was collected by a fraction collector and UO2þ 2 was measured by the Microwave Plasma-Atomic Emission Spectrophotometer and MRS-1 B. cereus was quantified by a UVevis spectrometry. 2þ To study the impact of phytate on UO2þ 2 immobilization, UO2 (10 mg/l), MRS-1 B. cereus (5 108 cells/ml) and phytate (10 mM) were introduced in the column. The column was left untouched for 10 days to allow MRS-1 B. cereus to release phosphate from organophosphates and interact with UO2þ 2 . 10 days later, the column was flushed with DI water first. When background UO2þ was 2 efluted from the column, the column was then flushed with MRS-1 B cereus suspension. DI water flushing was to investigate the possible mobilization of UO2þ by groundwater flow alone. The 2 MRS-1 B cereus flushing was to further investigate the mobilization of UO2þ 2 with the aid of mobile MRS-1 B cereus. Parallel column experiments in the absence of phytate were conducted as the controls. Above experiments were conducted at pH 4.2, 6.8 and 8.2, respectively with pH adjusted by 0.1 M HCl or 0.1 M NaOH. For each series of the column experiments, a fresh column was packed and prior to starting each experiment. The column had a volume of 31.4 ml with a pore volume of 17.3 ml. 173 ml (i.e., 10 pore volumes) of DI water was eluted through the column to stabilize the column. A conservative pulse tracer (chloride) breakthrough curve was generated separately to characterize the soil. UO2þ 2 transport in the SRS soil was described by the equilibriumkinetic two-site model (Toride et al., 1995):
bR
vC1 1 v2 C1 vC1 ¼ $ 2 uðC1 C2 Þ m1 C1 P vZ vT vZ
ð1 bÞR
vC2 ¼ uðC1 C2 Þ m2 C2 vT
where C1 and C2 are the dimensionless UO2þ 2 concentration in the solution and on the soil surface, respectively; b is the partition coefficient; R is the retardation factor (R ¼ 1 þ rbqKd , where rb is the soil bulk density, Kd is the partition coefficient of UO2þ 2 between the
353
solution and the soil, and q is the soil porosity); T is the dimensionless time; P is the Peclet number (P ¼ vL D , where v is the interstitial pore-water velocity, L is the length of the column, and D is the dispersion coefficient); Z is the dimensionless axial coordinate; u is the dimensionless mass transfer coefficient; and m1 and m2 are the dimensionless deposition coefficient in the solution and on the soil surface, respectively. For this research, it was assumed that the dominating UO2þ 2 deposition occurred from the solution. Subsequently, m2 was set to 0. For UO2þ transport, a pulse-type 2 boundary condition was used for the upper boundary and a zero gradient was assumed for the lower boundary (Reed, 1965):
qD
vC1 þ qC1 vZ
¼
Z¼0
qC0 0 < t t0 0 t > t0
vC1 ¼0 vZ Z¼L where q is the specific discharge (Darcian fluid flux), t0 is the duration of injection, and C0 is the initial UO2þ concentration. 2 Transport parameters in above equations were obtained by fitting the experimentally obtained UO2þ breakthrough data using an 2 implicit, finite-difference scheme of CXTFIT (Toride et al., 1995). All the parameters were optimized by minimizing the sum of squared differences between observed and fitted concentrations using the nonlinear least-square method (Toride et al., 1995). The simulated transport parameters are summarized in Table 1. It should be noted that above parameters were in the unit-less format. For deposition coefficient, m1, the corresponding values reported later were 1.5 103 min1, 9.2 103 min1 and 1.2 102 min1 for Scenario 1, 2 and 3, respectively. Effect of bacteria cannot be directly reflected by these models. On the other hand, the facilitated transport was reflected indirectly by the comparison of the simulated transport results with or without the aid of the facilitating bacteria.
3. Results and discussion Tracer (Cl) transport was studied before MRS-1 B. cereusfacilitated UO2þ 2 transport experiments in the SRS soil. Nearly all the input tracer was eluted from the column (Fig. 1). The tracer breakthrough curve was characterized by a breakthrough front and an elution tail and was simulated against transport equations. During the model simulation, the retardation factor was set to 1.0, i.e., Kd ¼ 0. This was based on the consideration that the tracer should not be retarded in the soil as the tracer was assumed not to adsorb in the soil. In addition, the deposition coefficient was set to zero, i.e., no retention of the tracer in the soil. This was true since nearly all the inputted tracer was eluted from the column at the end of the transport experiments. After the simulation, D was determined to be 1.30 cm2/min and then used for the simulations of UO2þ 2 transport. Breakthrough curves of MRS-1 B. cereus-facilitated UO2þ 2 transport in the SRS soil were characterized by a self-sharpening front, which became broader and diffuser at the elution limb (Fig. 1). The
Table 1 UO2þ 2 transport parameters simulated by the equilibrium-kinetic two-site model.
Chloride Scenario 1 Scenario 2 Scenario 3
b
u
m1
m2
0.990 ± 0.11 0.992 ± 0.08 0.998 ± 0.06 0.987 ± 0.12
0.02 ± 0.04 0.08 ± 0.05 1.43 ± 0.45 1.75 ± 0.61
Set to 0.00 5.2 105 ± 1.0 105 3.1 104 ± 1.5 105 4.2 104 ± 2.7 105
Set Set Set Set
to to to to
0.00 0.00 0.00 0.00
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2þ Fig. 1. UO2þ 2 breakthrough curves for different UO2 introduction scenarios.
long-lasting tails of the breakthrough curves indicated kineticcontrolled UO2þ 2 retention in the column. Under saturated conditions, UO2þ transport is controlled by kinetic adsorption and 2 equilibrium adsorption processes. The irreversible adsorption or kinetic adsorption in the SRS soil was evidenced by the reduced mass recovery (the integration of the breakthrough curve) during the transport and the equilibrium adsorption was demonstrated by delayed breakthrough reflected by the retardation factor. The UO2þ 2 breakthrough curves were simulated against the transport equations by CXTFIT (Toride et al., 1995). UO2þ 2 breakthrough curves were fitted well with the model and the accuracy of transport modeling was expressed by the sum of the squared differences between observed and fitted data. The mean square for error was in the range from 2.97 104 to 1.21 102. With the aid of MRS-1 B. cereus, UO2þ 2 transport was significantly facilitated by increased mass recovery of UO2þ 2 that passed through the column. Among the three scenarios, Scenario 1, i.e.,
UO2þ 2 pre-equilibriumed with MRS-1 B. cereus had the most facilitated transport observation with greatest UO2þ 2 mass recovery as compared to that of UO2þ 2 transport alone. For the other two scenarios, although increased mass recovery of UO2þ 2 was observed with the aid of the mobile MRS-1 B. cereus, the increase was much less pronounced. UO2þ 2 retention in the SRS soil was described by the deposition coefficient. With the aid of MRS-1 B. cereus, the corresponding deposition coefficient values decreased from 2 5.3 102 min1 for UO2þ min1, 2 transport alone to 1.2 10 9.2 103 min1 and 1.5 103 min1 for Scenario 3, 2 and 1, respectively. These deposition coefficient values reflected the irreversible adsorption of UO2þ 2 to the SRS soil and inversely corresponded to mass recovery of UO2þ 2 that passed through the column. The different deposition coefficient values of UO2þ 2 in the SRS soil for different facilitation scenarios indicated that the UO2þ in2 teractions with MRS-1 B. cereus affected the facilitated transport of UO2þ 2 . By comparing the breakthrough curves, it was discovered that for the three scenarios, UO2þ 2 breakthrough coincided with MRS-1 B. cereus transport (Fig. 2). This evidenced that UO2þ 2 transport was facilitated by MRS-1 B. cereus. The different bacterial transport observations among different scenarios were attributed to the formation of UO2þ 2 -bacteria complexation. For different scenarios, the different formation of UO2þ 2 -bacteria complexation impacted bacterial surface properties and the subsequently bacterial transport. The formation of different UO2þ 2 -bacteria complexation also affected UO2þ 2 transport. For Scenario 1, equilibrium was reached and the most complete UO2þ 2 -bacteria complexation was achieved, leading the most significant facilitated UO2þ 2 transport. For Scenario 2 and Scenario 3, less complete UO2þ 2 -bacteria complexation formation led to less significant facilitated UO2þ 2 transport. For Scenario 1 and Scenario 2, UO2þ 2 was carried by MRS-1 B. cereus during the transport in the soil, which also suffered from desorption and re-adsorption interactions between MRS-1 B. cereus and the soil. However, for Scenario 3, UO2þ 2 was desorbed from the SRS soil first and then transported with facilitating MRS-1 B. cereus. The ratio of UO2þ 2 /MRS-1 B. cereus eluted from the column as a function of pore volume clearly demonstrated different facilitated UO2þ 2
Fig. 2. UO2þ 2 /MRS-1 B. cereus. Ratio when eluted from the column.
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observations for different UO2þ introduction scenarios (Fig. 2). 2 Among these three scenarios, Scenario 1, i.e., UO2þ pre2 equilibriumed with MRS-1 B. cereus had the highest ratio, Scenario 3, i.e., UO2þ 2 pre-deposited in the soil column had the lowest ratio, and Scenario 2, i.e., UO2þ 2 introduced spontaneously with MRS-1 B. cereus was in between. When transported in the SRS soil, UO2þ 2 was also retarded as evidenced by the delayed breakthrough front, which was described by the retardation factor. Theoretically, retardation can be estimated from the breakthrough curve by (Kim and Kim, 2007; Li et al., 2009):
ð∞
C tdt tp C R ¼ ð0∞ 0 C 2 dt C 0 0 where t is the pore volume and tp is pulse width or pulse duration in pore volumes as defined by Dohse and Lion (1994). MRS-1 B. cereus facilitation also decreased UO2þ 2 retardation in the SRS soil. Retardation factor is the indication of the relative velocity of 2þ UO2þ 2 to that of water. UO2 transport was facilitated by the mobile bacteria, which was reflected by the retardation factor. The retardation factors obtained based on above equation were 4.04, 4.48 and 6.29 for Scenario 1, 2 and 3, respectively, which meant that water transport velocity was 4.04, 4.48 and 6.29 times that of UO2þ 2 . Above observations indicated that UO2þ 2 transport was most facilitated for Scenario 1 and least facilitated for Scenario 3. For the study of impact of phytate on UO2þ 2 immobilization, UO2þ and phytate were pre-deposited in the soil column first 2 before flushing with DI water and MRS-1 B. cereus suspension. The introduction of UO2þ 2 under different pH conditions was consistent with UO2þ 2 breakthrough curves almost overlapped with those of the control experiments (Fig. 3). Based on the mass balance, 0.038 mg of UO2þ 2 was pre-deposited in 8.09 g soil in the column. After 10 days, the column was flushed with DI water, followed by MRS-1 B. cereus suspension. During the 10 days’ period of time, MRS-1 B. cereus decomposed phytate to release phosphate. With the available phosphate and possible calcium, variable unsolvable UO-2-PO34 complexes and autunite [Ca(UO2)2(PO4)2] precipitation could be formed. The un-precipitated portion of UO2þ 2 was also adsorbed to the negative charged soil surface. Subsequently, minimal UO2þ 2 was flushed out by DI water (Fig. 3). However, when flushed by MRS-1 B. cereus suspension, the adsorbed UO2þ 2 may desorb from the soil and carried by MRS-1 B. cereus to break through the column. In addition, the UO-2-PO34 complexes and Ca(UO2)2(PO4)2 may also be possible to be carried out by mobile MRS-1 B. cereus. In another research, 10 mM organophosphate amendment was used between pH 4.5 and 7.0 and 73e95% of total uranium was precipitated after 120 h of incubation. This research demonstrated that indigenous bacterial within the uraniumcontaminated environment can hydrolyze organophosphate, especially under low pH conditions to immobilize UO2þ 2 (Beazley et al., 2007). The results were consistent with our observation that minimal UO2þ 2 was flushed out by DI water. Research using glycerol phosphate had similar observations that indigenous soil bacteria promoted the precipitation of UO2þ 2 at pH 5.5 and 7.0 (Beazley et al., 2011; Martinez et al., 2014; Salome et al., 2013). Phytate played an important role for the MRS-1 B. cereus-facil2þ itated UO2þ 2 transport as evidenced by the fact that much less UO2 was able to be flushed out by MRS-1 B. cereus with the addition of phytate as compared to that of the control. Facilitated UO2þ 2 transport by MRS-1 B. cereus with phytate addition was also a function of solution pH. With the increase of pH, less UO2þ 2 was flushed out by MRS-1 B. cereus (Fig. 3). To understand the impact of
Fig. 3. UO2þ 2 breakthrough curves by DI water and MRS-1 B. cereus. Suspension after 10 days of incubation with phytate.
pH on MRS-1 B. cereus-facilitated UO2þ transport with phytate 2 addition, the graph of uranium solubility versus pH was con4þ structed based on the introduced UO2þ 2 , released PO4 and pore water calcium concentration. Uranium solubility decreased with the increase of pH until pH 8, and then increased with the continuous increase of pH (Fig. 4). Because of the high solubility at pH 4, UO2þ 2 was easier to be transported with the facilitation of MRS-1 B. cereus. On the contrary, due to the low solubility at pH 8.2, less UO2þ was flushed out resulting from UO2þ precipitation. 2 2 Assuming a first order release from the SRS soil (Mishra et al., 2012; Rahmani-Sani et al., 2015), the release coefficient linearly decreased with the increase of pH (Fig. 5).
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Fig. 4. UO2þ 2 Speciation as a Function of pH.
Fig. 5. UO2þ 2 Release Coefficient as a Function of pH.
UO2þ 2 adsorption in the SRS soil was estimated by integrating the diffuse front of the breakthrough curves (Burgisser et al., 1993) (Fig. 6). Based on this method, the concentration of equilibrium adsorbed UO2þ 2 in the SRS soil, S, can be obtained by integrating the experimental record of the retention time t(c) if the dispersion term can be neglected (D ¼ 0): Fig. 6. UO2þ 2 sorption isotherms in SRS soil.
S¼
q rb ð1 qÞ
ðC 0
ðC tðC 0 Þ 0 1 dC tðC 0 ÞKc dC 0 t0 0
where t0 ¼ Lv and is the average UO2þ 2 travel time in the column (min). The insignificant role of hydrodynamic dispersion on UO2þ 2 transport has been proven to be valid (Unice and Logan, 2000). In their research, they demonstrated that hydrodynamic dispersion can be neglected for Pe > 100. For this research, all the column experiments were performed with Pe > 100, thus above equation can be used to determine the UO2þ 2 adsorption isotherms. The adsorption of UO2þ 2 in the SRS soil increased with the increase of the equilibrium concentration. UO2þ had greater 2 adsorption in the SRS soil with the phytate addition than those of without, implying UO-2-PO34 complexes and/or [Ca(UO2)2(PO4)2] precipitation occurred, which increased its sorption in the SRS soil.
UO2þ 2 adsorption in the SRS soil was described by Langmuir isotherms (Adachi et al., 1997; Hajratwala, 1982):
S¼
qmax KCe 1 þ KCe
where S is the adsorbed UO2þ 2 in the SRS soil; qmax is the maximum mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface, Ce is the UO2þ 2 equilibrium concentration, and K is the Langmuir adsorption constant, which increases with the increase of the binding energy of adsorption. UO2þ 2 had qmax values of 0.012, 0.015 and 0.020 mM/g for pH 4.2, 6.8 and 8.2 without the addition of phytate. With the addition phytate, these values increased to 0.020, 0.028 and 0.038 mM/g, respectively. qmax is the maximum
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mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface. Above observations indicated that the increase of pH increased the mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface. In the presence of phytate, the mass of UO2þ 2 required to form a mono-layer coverage of the SRS soil surface further increased. The underlying principle behind these observations was the formation of bonding between UO2þ 2 and adsorption receptor sites on SRS soil surfaces. Increase of pH and the presence of phytate increased the bonding energy between UO2þ 2 and the adsorption receptor sites on SRS soil surfaces. 4. Conclusions Under oxic conditions at the uranium-contaminated sites, uranium is primarily in dissolved form of UO2þ 2 , which can get equilibrium with indigenous MRS-1 B. cereus. The subsequent MRS-1 B. cereus transport can facilitate UO2þ 2 transport. This facilitated transport may also occur before equilibrium is reached. In addition, UO2þ 2 can be adsorbed to the soil before MRS-1 B. cereus transport occurs. These three possible scenarios of MRS-1 B. cereus-facilitated UO2þ transport were investigated in this research. In practice, 2 addition of phytate has been practiced as a means of UO2þ 2 immobilization. The presence of phytate on MRS-1 B. cereus-facilitated UO2þ 2 transport was also explored. In this research, transport of UO2þ 2 in the SRS soil was found to be significantly facilitated by MRS-1 B. cereus. The facilitation results were controlled by UO2þ 2 interactions with MRS-1 B. cereus. Pre-equilibrium with MRS-1 B. cereus before the introduction had the most facilitated transport observations. Further analysis showed that MRS-1 B. cereusfacilitated transport was due to the adsorption of UO2þ 2 on the MRS1 B. cereus surface based on the quantification of the ratio of UO2þ 2 / MRS-1 B. cereus eluted from the column. Phytate played an important role for the MRS-1 B. cereus-facilitated UO2þ 2 transport, which was found to be a function of solution pH. The immobilization of UO2þ 2 with the addition of phytate increased with the increase of pH within the pH range of this study because uranium solubility decreased with the increase of pH until pH 8. Acknowledgements The work was supported by Department of Energy Minority Serving Institution Partnership Program (MSIPP) managed by the Savannah River National Laboratory under SRNS contract DE-AC0908SR22470. Appendix A. Supplementary data Supplementary data related to this article can be found at https://doi.org/10.1016/j.chemosphere.2019.02.064. References Adachi, S., Panintrarux, C., Matsuno, R., 1997. Methods for estimating the parameters of nonlinear adsorption isotherms of Langmuir and Freundlich types from a response curve of pulse input of an adsorbate. Biosci. Biotechnol. Biochem. 61, 1626e1633. Arey, J.S., Seaman, J.C., Bertsch, P.M., 1999. Immobilization of uranium in contaminated sediments by hydroxyapatite addition. Environ. Sci. Technol. 33, 337e342. Beazley, M.J., Martinez, R.J., Sobecky, P.A., Webb, S.M., Taillefert, M., 2007. Uranium biomineralization as a result of bacterial phosphatase activity: insights from bacterial isolates from a contaminated subsurface. Environ. Sci. Technol. 41, 5701e5707. Beazley, M.J., Martinez, R.J., Webb, S.M., Sobecky, P.A., Taillefert, M., 2011. The effect of pH and natural microbial phosphatase activity on the speciation of uranium in subsurface soils. Geochem. Cosmochim. Acta 75, 5648e5663. Buhl, M., Diss, R., Wipff, G., 2005. Coordination environment of aqueous uranyl (VI)
357
ion. J. Am. Chem. Soc. 127, 13506e13507. Burgisser, C.S., Cernik, M., Borkovec, M., Sticher, H., 1993. Determination of nonlinear adsorption-isotherms from column experiments: an alternative to batch studies. Environ. Sci. Technol. 27, 943e948. Dohse, D.M., Lion, L.W., 1994. Effect of microbial polymers on the sorption and transport of phenanthrene in a low-carbon sand. Environ. Sci. Technol. 28, 541e548. Dong, W., Tokunaga, T.K., Davis, J.A., Wan, J., 2012. Uranium (VI) adsorption and surface complexation modeling onto background sediments from the F-area Savannah River Site. Environ. Sci. Technol. 46, 1565e1571. Fortin, C., Dutel, L., Garnier-Laplace, J., 2004. Uranium complexation and uptake by a green alga in relation to chemical speciation: the importance of the free uranyl ion. Environ. Toxicol. Chem. 23, 974e981. Hajratwala, B.R., 1982. Potential errors in determining Freundlich and Langmuir constants from adsorption isotherms. J. Pharmacol. Sci. 71, 125e126. Horii, S., Matsuno, T., Tagomori, J., Mukai, M., Adhikari, D., Kubo, M., 2013. Isolation and identification of phytate-degrading bacteria and their contribution to phytate mineralization in soil. J. Gen. Appl. Microbiol. 59, 353e360. Huynh, T.S., Vidaud, C., Hagege, A., 2016. Investigation of uranium interactions with calcium phosphate-binding proteins using ICP/MS and CE-ICP/MS. Metall 8, 1185e1192. Ibeanusi, V., Pathak, A., Chauhan, A., Hoyle-Gardner, J., Cooper, T., Turker, L., Howard, H., Obinegbo, O., Chen, G., Seaman, J., 2018. Genome-centric evaluation of Bacillus sp. strain: ATCC55673 and response to uranium biomineralization. Significance Bioeng Biosci 2, 1e8. Kelly, S.D., Boyanov, M.I., Bunker, B.A., Fein, J.B., Fowle, D.A., Yee, N., Kemner, K.M., 2001. XAFS determination of the bacterial cell wall functional groups responsible for complexation of Cd and U as a function of pH. J. Synchrotron Radiat. 8, 946e948. Kim, M., Kim, S.B., 2007. Modeling contaminant transport in a three-phase groundwater system with the Freundlich-type retardation factor. Environ. Technol. 28, 205e216. Li, D., Seaman, J.C., Chang, H.S., Jaffe, P.R., Koster van Groos, P., Jiang, D.T., Chen, N., Lin, J., Arthur, Z., Pan, Y., Scheckel, K.G., Newville, M., Lanzirotti, A., Kaplan, D.I., 2014. Retention and chemical speciation of uranium in an oxidized wetland sediment from the Savannah River Site. J. Environ. Radioact. 131, 40e46. Li, M.H., Wang, T.H., Teng, S.P., 2009. Experimental and numerical investigations of effect of column length on retardation factor determination: a case study of cesium transport in crushed granite. J. Hazard Mater. 162, 530e535. Lim, B.L., Yeung, P., Cheng, C., Hill, J.E., 2007. Distribution and diversity of phytatemineralizing bacteria. ISME J. 1, 321e330. Liu, C., Gorby, Y.A., Zachara, J.M., Fredrickson, J.K., Brown, C.F., 2002. Reduction kinetics of Fe(III), Co(III), U(VI), Cr(VI), and Tc(VII) in cultures of dissimilatory metal-reducing bacteria. Biotechnol. Bioeng. 80, 637e649. Martinez, R.J., Wu, C.H., Beazley, M.J., Andersen, G.L., Conrad, M.E., Hazen, T.C., Taillefert, M., Sobecky, P.A., 2014. Microbial community responses to organophosphate substrate additions in contaminated subsurface sediments. PLoS One 9, e100383. Mishra, S., Maity, S., Bhalke, S., Pandit, G.G., Puranik, V.D., Kushwaha, H.S., 2012. Thermodynamic and kinetic investigations of uranium adsorption on soil. J. Radioanal. Nucl. Chem. 294, 97e102. Rahmani-Sani, A., Hosseini-Bandegharaei, A., Hosseini, S.H., Kharghani, K., Zarei, H., Rastegar, A., 2015. Kinetic, equilibrium and thermodynamic studies on sorption of uranium and thorium from aqueous solutions by a selective impregnated resin containing carminic acid. J. Hazard Mater. 286, 152e163. Reed, K.W., 1965. A time dependent boundary condition for neutron transport equation. Am. Math. Mon. 72, 454. Roden, E.E., Scheibe, T.D., 2005. Conceptual and numerical model of uranium(VI) reductive immobilization in fractured subsurface sediments. Chemosphere 59, 617e628. Salome, K.R., Green, S.J., Beazley, M.J., Webb, S.M., Kostka, J.E., Taillefert, M., 2013. The role of anaerobic respiration in the immobilization of uranium through biomineralization of phosphate minerals. Geochem. Cosmochim. Acta 106, 344e363. Seaman, J.C., Hutchison, J.M., Jackson, B.P., Vulava, V.M., 2003. In situ treatment of metals in contaminated soils with phytate. J. Environ. Qual. 32, 153e161. Suleimanova, A.D., Beinhauer, A., Valeeva, L.R., Chastukhina, I.B., Balaban, N.P., Shakirov, E.V., Greiner, R., Sharipova, M.R., 2015. Novel glucose-1-phosphatase with high phytase activity and unusual metal ion activation from soil bacterium Pantoea sp. Strain 3.5.1. Appl. Environ. Microbiol. 81, 6790e6799. Toride, N.L., Leij, F.J., van Genuchten, M. Th., 1995. The CXTFIT Code for Estimating Transport Parameters from Laboratory or Field Experiments, Version 2.1. U.S. Salinity Laboratory, Riverside, CA. Truex, M.J., Peyton, B.M., Valentine, N.B., Gorby, Y.A., 1997. Kinetics of U (VI) reduction by a dissimilatory Fe (III)-reducing bacterium under non-growth conditions. Biotechnol. Bioeng. 55, 490e496. Unice, K.M., Logan, B.E., 2000. Insignificant Role of hydrodynamic dispersion on bacterial transport. J. Environ. Eng. 126, 491e500. Vazquez, G.J., Dodge, C.J., Francis, A.J., 2007. Interactions of uranium with polyphosphate. Chemosphere 70, 263e269. Yeomans, J.C., Bremner, J.M., 1998. A rapid and precise method for routine determination of organic carbon in soil. Commun. Soil Sci. Plant Anal. 19, 1467e1476.