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Immobilization of U(VI) from oxic groundwater by Hanford300 Area sediments and effects of Columbia River water

Bulbul Ahmed a, Bin Cao a,b,1, Bhoopesh Mishra c, Maxim I. Boyanov c,Kenneth M. Kemner c, Jim K. Fredrickson b, Haluk Beyenal a,*aThe Gene and Linda Voiland School of Chemical Engineering and Bioengineering, Washington State University, Spokane St, PO Box 642710,

Pullman, WA 99164-2710, United Statesb Pacific Northwest National Laboratory, Richland, WA, United StatescArgonne National Laboratory, Argonne, IL, United States

a r t i c l e i n f o

Article history:

Received 15 December 2011

Received in revised form

8 May 2012

Accepted 15 May 2012

Available online 23 May 2012

Keywords:

Uranium immobilization

Bioremediation

Biofilm

Groundwater

Columbia River water

Remobilization

* Corresponding author. Tel.: þ1 509 335 660E-mail address: beyenal@wsu.edu (H. Bey

1 Present address: School of Civil & EnvNanyang Technological University, Singapor0043-1354/$ e see front matter ª 2012 Elsevdoi:10.1016/j.watres.2012.05.027

a b s t r a c t

Regions within the U.S. Department of Energy Hanford 300 Area (300 A) site experience

periodic hydrologic influences from the nearby Columbia River as a result of changing river

stage, which causes changes in groundwater elevation, flow direction and water chemistry.

An important question is the extent to which the mixing of Columbia River water and

groundwater impacts the speciation and mobility of uranium (U). In this study, we

designed experiments to mimic interactions among U, oxic groundwater or Columbia River

water, and 300 A sediments in the subsurface environment of Hanford 300 A. The goals

were to investigate mechanisms of: 1) U immobilization in 300 A sediments under bulk oxic

conditions and 2) U remobilization from U-immobilized 300 A sediments exposed to oxic

Columbia River water. Initially, 300 A sediments in column reactors were fed with U(VI)-

containing oxic 1) synthetic groundwater (SGW), 2) organic-amended SGW (OA-SGW),

and 3) de-ionized (DI) water to investigate U immobilization processes. After that, the

sediments were exposed to oxic Columbia River water for U remobilization studies. The

results reveal that U was immobilized by 300 A sediments predominantly through reduc-

tion (80e85%) when the column reactor was fed with oxic OA-SGW. However, U was

immobilized by 300 A sediments through adsorption (100%) when the column reactors

were fed with oxic SGW or DI water. The reduced U in the 300 A sediments fed with OA-

SGW was relatively resistant to remobilization by oxic Columbia River water. Oxic

Columbia River water resulted in U remobilization (w7%) through desorption, and most of

the U that remained in the 300 A sediments fed with OA-SGW (w93%) was in the form of

uraninite nanoparticles. These results reveal that: 1) the reductive immobilization of U

through OA-SGW stimulation of indigenous 300 A sediment microorganisms may be viable

in the relatively oxic Hanford 300 A subsurface environments and 2) with the intrusion of

Columbia River water, desorption may be the primary process resulting in U remobilization

from OA-SGW-stimulated 300 A sediments at the subsurface of the Hanford 300 A site.

ª 2012 Elsevier Ltd. All rights reserved.

7; fax: þ1 509 335 4806.enal).ironmental Engineering,e.ier Ltd. All rights reserved

Singapore Centre on Environmental Life Sciences Engineering,

.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 83990

1. Introduction

Because of historical nuclear materials processing, storage and

waste disposal facilities, soils, sediments, and groundwater at

DOE sites such as Hanford have been contaminated with

uranium (U) (Palmisano and Hazen, 2003; Riley and Zachara,

1992). Although several thousand tons of sediment have been

excavated and removed from the Hanford 300 Area (300 A),

significant U contamination persists in the groundwater

beneath the 300 A site and there is a risk of it migrating to the

Columbia River (Brown et al., 2008; McKinley et al., 2007;

Peterson, 2001). Furthermore, the Hanford 300 A subsurface is

subject tohydrologicfluctuationsassociatedwithchanges inthe

stage of the ColumbiaRiver. These fluctuations include changes

in groundwater elevation and flow direction (Brown et al., 2008;

Peterson, 2001). Thus, the area is highly subject to U remobili-

zation due to changes in water chemistry attributed to periodic

saturation of the base region of the vadose zone by changes in

the water levels of the Columbia River (Brown et al., 2008;

Peterson et al., 2008; Yabusaki et al., 2008; Zachara et al., 2005).

These mixing events can potentially transport U contaminants

and pose a serious threat to the natural aquatic environment of

the Columbia River and to human health (Peterson, 2001).

Natural attenuation, the strategy of harnessing indigenous

microbial processes for cleanup, has been extensively studied

in relation to U remediation (Abdelouas et al., 1998; Anderson

et al., 2003; Istok et al., 2004; North et al., 2004; Senko et al.,

2002). The approach is to promote the reduction of soluble

U(VI) to sparingly soluble U(IV) species through indigenous

microbial activities, thereby preventing U migration in the

subsurface environment (Anderson et al., 2003). Dissimilatory

metal-reducingmicroorganisms typically reduce U(VI) to form

nanocrystalline uraninite (UO2) in the periplasm and extra-

cellularmaterial (Burgos et al., 2008; Marshall et al., 2006, 2009;

Marsili et al., 2005; Suzuki et al., 2002). However, microbially

mediated non-uraninite reduction products have also been

reported (Bernier-Latmani et al., 2010; Boyanov et al., 2011;

Dalla Vecchia et al., 2010; Fletcher et al., 2010; Sivaswamy

et al., 2011). U(VI) reduction through the stimulation of

dissimilatory metal-reducing microbial activities has been

proposed for the in situ immobilization of U in the subsurface

(Lovley, 1993; Lovley et al., 1991; N’Guessan et al., 2008). The

presence of naturally occurring electron acceptors such as O2,

Fe(III) and Mn(IV) may impede U immobilization by buffering

against redox changes and/or reoxidizing U(IV) through

abiotic or indirect biotic reactions (Fredrickson et al., 2002;

Moon et al., 2009; Sani et al., 2005). To understand the effects

of alternative electron acceptors on the stability of bioreduced

U, column studies using sediments (Moon et al., 2009, 2007;

Tokunaga et al., 2008; Wan et al., 2005) and groundwater

(Moon et al., 2009; Tokunaga et al., 2008) from DOE contami-

nated sites have been carried out. In these studies, U was

reduced through biostimulation under anaerobic conditions

and then various oxidants were introduced to determine their

effects on U remobilization (Moon et al., 2009, 2007; Tokunaga

et al., 2008). Moon et al. (2009) showed that biogenic iron

sulfide precipitates are more effective at protecting biogenic

U(IV) in the presence of O2 and nitrate. In another study, Moon

et al. (2007) showed that acetate stimulation of sediment from

Old Rifle, CO resulted in the immobilization of 62e92% of the

influent U under anaerobic conditions. However, once the

supply of acetate stopped, the presence of dissolved oxygen

(DO) (0.27 mM) or nitrate (1.6 mM) in the influent medium

caused the reoxidation of 88% or 97% of the immobilized U,

respectively. Takunaga et al. (2008) reported changes in the

oxidation states of U, Fe and Mn in U-contaminated sediment

from Oak Ridge National Laboratory, TN and concluded that

Fe(III) was responsible for U(IV) reoxidation. These studies

demonstrated that U immobilized in biostimulated sediment

can be remobilized from the sediment when exposed to

electron acceptors such as O2, nitrate and Fe(III). However,

data generated in different sediment experimentsmay lead to

different conclusions about the stability of bioreduced U

exposed to O2, depending on sediment microbial, chemical

and physical properties.

A high DO concentration usually leads to U remobilization

through reoxidation (Moon et al., 2007). However, microorgan-

isms in natural environments (e.g., wateresediment interfaces

in the subsurface) are commonly found in the form of biofilms

(Decho, 2000; Jones et al., 2010; Weidler et al., 2007), and the

heterogeneity of geochemical gradients in biofilms in subsur-

face sediments (also known as sediment biofilms) may play an

important role in controlling U mobility (Nguyen et al., 2012).

Based on previous studies on microenvironments in model

biofilms (Lewandowski and Beyenal, 2007; McLean et al., 2008;

Nguyen et al., 2012), we hypothesize that, under bulk aerobic

conditions,withO2 limitedorevendepleted inside thesediment

biofilms because of microbial O2 consumption and diffusion

limitation, U(VI) can be immobilized within the sediment bio-

films as U(IV) and remain resistant to the O2 present in the bulk

aqueous phase as long as the active biofilms are maintained.

Since the groundwater in the Hanford 300 A site has been

reported to be at or near saturation with O2 throughout the

year (DO up to 8.22 � 0.26 mg/L) (Zachara, 2009), it is crucial to

understand U mobility in the Hanford 300 A site in the pres-

ence of oxic groundwater. In addition, because of the periodic

hydrologic mixing of Columbia River water with groundwater,

knowledge of the effects of Columbia River water on U

mobility is also crucial to a comprehensive understanding of

the fate of U in the subsurface at this site. Therefore, the goals

of this study were to investigate mechanisms of (1) U immo-

bilization in 300 A sediments in synthetic groundwater (SGW)

column reactors and (2) U remobilization from U-immobilized

300A sediments under bulk oxic conditions by oxic Columbia

Riverwater. X-ray absorption near edge structure (XANES) and

extended X-ray absorption fine structure (EXAFS) spectros-

copies were used to characterize U oxidation states, elucidate

the local coordination environment around solid-phase-

associated U atoms and to determine the average particle

sizes of reduced uraninite species in the 300A sediments.

2. Materials and methods

2.1. Sediment and groundwater characteristics

A core sample (well ID# C-6190, 2-13/C/330-340, Hanford

formation) was collected using resonant sonic drilling from

the DOE Hanford 300 A, Richland, WA (http://ifchanford.pnl.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 8 3991

gov). A site map showing the Hanford 300A and the Columbia

River and the conceptual model for the interaction between

300A groundwater and Columbia River water are shown in

Fig. S1. The lithology of the Hanford formation within 300A is

a gravel-dominated, poorly sorted mixture with lesser

amounts of sand and silt (Bjornstad and Horner, 2008;

Bjornstad et al., 2009). The groundwater in this well has a DO

content that varies with season from 5.9 mg/L to 8.3 mg/L

because of Columbia River water infiltration (Zachara, 2009).

2.2. Sediment column reactors

The 300 A sediments were prepared in column reactors. Three

column reactors were prepared by packing the sediments

from the core sample into polycarbonate tubes (internal

diameter of 2.54 cm and length of 40 cm) and fitting themwith

flow distributors. These were made of 0.3-cm glass beads

entrapped between two plastic sieves for uniform flow

distribution and placed in the inlet and outlet of each column

reactor (Marsili et al., 2007). A schematic illustration and

a picture of the 300 A sediments in column reactors are shown

in Supplemental Information (SI) Fig. S2. Each column reactor

contained 350 � 15 g of sediment, and the porosity of the

packed sediment was 38.7 � 5.0%. The column reactors were

fed with oxic SGW, organics-amended SGW (OA-SGW e 2 mM

lactate, 2 mM malate, 2 mM succinate and 2 mM fumarate in

SGW) or de-ionized (DI) water. Table S1 summarizes the

medium pH and DO used in the column reactors. The SGW

consisted of 7.0 mg/L KHCO3, 25.2 mg/L MgSO4, 42.4 mg/L

Ca(NO3)2$4H2O, 61.7 mg/L CaCl2$2H2O, 19.9 mg/L Na2SO4,

92.4mg/L NaHCO3 (Nguyen et al., 2012; Stoliker et al., 2011; Yin

et al., 2011), which was based on the groundwater composi-

tion at the Hanford 300 A site. Themediumflow for all column

reactors was in an upward direction and maintained at a flow

rate of 115 mL/day. The experimental steps during the period

of the study are given in Fig. S3. After a five-month operation

of these column reactors, the 300A sediments from each

column reactor were imaged and then the column reactors

were used for U immobilization for seven more months.

During the five-month operation, samples were collected

periodically from the effluent to measure soluble total iron

and manganese concentrations.

2.3. Scanning electron microscopy

After fivemonths of operation of the column reactors fed with

oxic OA-SGW, SGW, and DI water (Fig. S3), sediments from

each reactorwere imaged using scanning electronmicroscopy

(SEM e FEI 200F, FEI Company, Hillsboro, OR). Low-vacuum

SEM was used, and the images were taken directly using wet

samples with no dehydration or staining at an acceleration

voltage of 10 kV, a vacuum pressure of 120 Pa, and a spot size

of 3.5 nm (Cao et al., 2011a, 2011b).

2.4. U immobilization in 300 A sediments under bulkoxic conditions

The column reactors were used for U immobilization with the

same flow rate that was used during 300 A sediment prepa-

ration. The composition of the medium supplied to each

column did not change, except that uranyl chloride (UO2Cl2)

was amended to a final concentration of 126 mM. Samples

were taken periodically from the effluent, and concentrations

of soluble total iron, manganese, lactate and U were

determined.

2.5. U remobilization from 300 A sediments by oxicColumbia River water

U remobilization was studied using syringe columns (3 mL)

that were packed with 300 A sediments taken from the

column reactors, which had been exposed to U(VI) for seven

months (Fig. S3). A schematic illustration and a picture of the

syringe columns are shown in Fig. S4. Columbia River water

was collected from the river bank close to the Hanford 300 A

and kept in a bucket in darkness at room temperature

(22 � 2 �C) with less than 5% v/v air as headspace. The back-

ground U(VI) concentration in the river water was 0.004 mM

(�5%). After particles were removed through filtration (5 mm),

Columbia River water with a DO concentration of 7.6 �0.05 mg/L was continuously fed into the syringe columns at

a flow rate of 0.1 mL/h Table S1 lists the pH and DO of the

Columbia River water used in syringe columns for the U

remobilization experiments. The chemical composition of the

Columbia River water has previously been reported (Poston

et al., 2009; Yin et al., 2011). Effluent samples were collected

periodically to determine total U concentration. At the end of

the remobilization experiment, the DO concentrations at the

outlets of the syringe columns were measured using micro-

electrodes (Lewandowski and Beyenal, 2007) to assess the O2

consumption by the 300 A sediments, and U LIII-edge X-ray

absorption fine structure (XAFS) spectroscopy was used to

examine U speciation in the solids.

2.6. XAFS measurement and data analysis

U-immobilized 300 A sedimentswere collected from the inlets

and the outlets of the column reactors and kept in serum

bottles with an O2-free N2 headspace to maintain redox

integrity for XAFS analyses. Data collection and analysis were

performed as described previously (Boyanov et al., 2007;

Kemner and Kelly, 2007). The samples were mounted under

anoxic conditions inside a void cut in a plexiglass slide and

sealed with Kapton film windows. U LIII edge (17166-eV) XAFS

measurements were performed at sector 10-ID of the

Advanced Photon Source (Segre et al., 2000). The beamline

undulator was tapered, and the energy of the incident X rays

was scanned using a Si(111) cryogenically cooled double-

crystal monochromator. Harmonic content was removed by

reflection from a Rh-coated harmonic rejection mirror. XAFS

scans were collected in quick scanning mode (3 min each).

Energy calibration was maintained by the simultaneous

collection of data from a hydrogen uranyl phosphate stan-

dard. The final spectrum was produced by averaging 30 quick

XAFS scans. Processing of the raw data was done using

ATHENA (Ravel and Newville, 2005). The resulting c(k) k3

XAFS data are included in the Supplementary Information

section to illustrate the quality of the data (Fig. S5). The EXAFS

data were refined using theoretical models. The crystal

structure of uraninite, UO2, was used to generate theoretical

Fig. 1 e Total U immobilized in 300 A sediments in column

reactors fed with oxic OA-SGW (,), SGW (>) or DI water (D)

with U. The numbers (I, II, III, and IV) refer to the different U

immobilization phases, which have different U

immobilization rates.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 83992

EXAFS spectra with the programs ATOMS and FEFF8

(Ankudinov et al., 1998; Ravel and Newville, 2005; Wyckoff,

1960). Data were analyzed using the UWXAFS package (Stern

et al., 1995) implemented in ARTEMIS (Ravel and Newville,

2005). The Fourier-transformed c(R) spectra were simulta-

neously fitted at three different k-weights (k1, k2, k3).

Several U standards were prepared and analyzed with U LIIIedge XANES and EXAFS for spectral comparisons with data

from U in the sediments. The U(IV) and U(VI) standards for

comparing XANES data were biogenic nanoparticulate urani-

nite and hydrogen uranyl phosphate, respectively. In addition

to the XANES standards, the standards for EXAFS analysis

included data from an acidic uranyl chloride aqueous solution

(pH 3), U(VI) adsorbed to goethite (200 mM uranyl adsorbed to

1.5-g/L goethite at pH 7.5 in the presence of 500 mM NaHCO3),

and data from a fully reduced non-uraninite U(IV) species

(Boyanov et al., 2011).

2.7. Analytical methods

Total U concentration was analyzed using a KPA analyzer

(Chemchek Instruments, Richland,Washington). The samples

were processed in 20-mL glass scintillation vials following the

previously described method (Hedaya et al., 1997). Briefly,

a 200-ml unfiltered effluent sample was mixed with 2 mL of

16 M HNO3 and 0.5 mL of H2O2 and incubated in an oven near

100 �C. The heat was gradually increased to 200 �C until the

sample dried completely. The sample was then ashed at

550 �C in a muffle furnace for at least 2 h. Two mL of 0.82 M

HNO3 were added to each vial and mixed thoroughly before

U(VI) measurements were taken using KPA (Sani et al., 2002).

The soluble total iron and manganese concentrations were

determined using inductively coupled plasma-mass spec-

trometry (ICP-MS). The samples were filtered through 0.22-mm

syringe filters and acidified to keep the pH less than 2.0

(1.9 � 0.1). Lactate concentrations were analyzed using high-

performance liquid chromatography (HPLC e Aminex HPX-

87H column, HP 1100 HPLC system).

3. Results and discussion

3.1. U immobilization in Hanford 300 A sedimentsunder bulk oxic conditions

Fig. 1 shows total U immobilized in 300 A sediments in column

reactors fed with oxic OA-SGW, SGW, or DI water containing

126 mM U(VI). We divided the U immobilization period into

four different phases, asmarked in Fig. 1, based on differences

in U immobilization rates. Table 1 summarizes the U immo-

bilization rate during each phase, calculated from Fig. 1, and

the oxidation states of the immobilized U, estimated from the

XANES analyses of the sediments collected at the end of the

experiment. Over the entire operating period (w219 days) U

immobilization was highest in the 300 A sediments in the

column reactor fed with oxic DI water. For the first 150 days,

the total U immobilization was higher in the reactors fed with

SGW than in the reactor fed with OA-SGW. However, during

the last w75 days of the experiment, the total U immobiliza-

tion rate in the reactor fed with OA-SGW increased and was

higher than the rate in the reactor fed with SGW. Represen-

tative SEM images of 300 A sediments from the column reac-

tors fed with oxic OA-SGW, SGW, or DI water are shown in

Fig. S6. Cells were not observed by SEM in 300 A sediments

from the column reactor fed with oxic DI water but were

common in sediments from column reactors fed with oxic

SGW or OA-SGW. The 300 A sediments from the column

reactor fed with oxic OA-SGW also had extracellular poly-

meric substances (EPS), observed in Fig. S6. Synchrotron-

based U LIII edge XANES and EXAFS measurements taken at

the end of the experiments for each column reactor to deter-

mine the oxidation state of the immobilized U and to elucidate

the local coordination environment around the U atoms are

shown in Fig. 2.

The total U immobilization ratewasw3.4mg/day in phases

I & II andw2.39mg/day in phases III & IV in 300 A sediments in

the column reactor fed with DI water. The difference in total U

immobilization rate between the 300 A sediments fed with

SGW or OA-SGW and those fed with DI water could be due to

differences in: (1) U(VI) speciation (e.g., U(VI) complexation

with the Ca2þ and HCO�3 present in SGW) or (2) the sorption

characteristics of the U(VI) complex in sediments in a column

reactor (Stewart et al., 2010; Wazne et al., 2003). The major

U(VI) species present in DI water are U(VI) hydroxide (2.51%

UO2(OH)2, 0.64% UO2OHþ, 1.32% ðUO2Þ3ðOHÞ�7 , 59.04%

ðUO2Þ3ðOHÞþ5 , and 35.83% ðUO2Þ3ðOHÞþ7 Þ, but in the presence of

carbonate and calcium in SGW or OA-SGW, aqueous uranyl-

carbonate (0.75% UO2ðCO3Þ�22 , 1.99% UO2ðCO3Þ�4

3 Þ and

calcium-uranyl-carbonate (41.54% CaUO2ðCO3Þ�23 , 55.67%

Ca2UO2(CO3)3) complexes are predominant (Table S3).

Carbonate forms a strong complex with U(VI), increasing

solubility and decreasing the propensity for uranium reten-

tion on mineral surface sites (e.g., iron hydroxide) (Wazne

et al., 2003). Uranyl carbonates and ternary uranyl-calcium-

carbonates have lower affinities for adsorbing to mineral

surface sites such as iron hydroxide, goethite, ferrihydrite,

quartz, soil or bacterial surfaces compared with uranyl

hydroxide complexes, which are natural retardants for U

Table 1 e U immobilization rates during each phase of the U immobilization experiment and oxidation states of theimmobilized U at the end of the U immobilization experiment for column reactors (% deviation in U immobilization ratewas w3%).

300 A sedimentsfed with

Phases

U immobilization rate (mg/day) U oxidation state

I II III IV

OA-SGW 2.23 1.45 1.55 3.20 80e85 % U(IV)

SGW 2.51 2.16 1.70 2.03 100% U(VI)

DI water 3.40 3.34 2.39 2.39 100% U(VI)

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 8 3993

migration (Brooks et al., 2003; Fox et al., 2006; Gorman-Lewis

et al., 2005; Stewart et al., 2010; Wazne et al., 2003; Zheng

et al., 2003). The U immobilization in 300 A sediments fed

with DI water is likely due to U adsorption to sediments, since

we did not add organic compounds to support microbial

growth and we did not observe attached cells on the surfaces

(Fig. S6). Synchrotron-based U LIII edge XANES analysis

showed that the U immobilized in 300 A sediments from the

column reactor fed with DI water was U(VI) (Table 1, Fig. 2b).

EXAFS analysis showed that the inlet sample from the column

reactor with 300 A sediments fed with DI water was a combi-

nation of aqueous uranyl and an inner-sphere complex of

uranyl adsorbed to metal oxides, while the outlet samples

were found to be predominantly the latter (Fig. 2d). These

Fig. 2 e U LIII edge XANES spectra of the inlet and outlet U-imm

(A) OA-SGW (B) SGW or DI water plotted with U(IV) and U(VI) st

the inlet and outlet U-immobilized sediments from (C) the colu

nanoparticulate, and (D) the column reactors fed with SGW or D

uranyl adsorbed to goethite.

conclusions are based on the EXAFS spectra from the sedi-

ments closely matching that of the goethite-sorbed U(VI)

standard, in which U(VI) has been shown previously to adsorb

as an inner-sphere complex (Redden et al., 2001).

There was no supply of exogenous organics to the column

reactor fed with SGW during the period of study. Although we

observed the presence of cells in 300 A sediments in the

column reactor fed with SGW, soluble iron and manganese

were not detected in the effluent before or during the U

immobilization experiments (Fig. S7), indicating that condi-

tions within the column were not strongly reducing. There-

fore, we expected that U immobilization in the column reactor

fed with SGW would be mostly through the adsorption of

U(VI). XANES and EXAFS analyses of U-immobilized 300 A

obilized sediments from column reactors fed with

andards. Fourier transforms of U LIII edge EXAFS spectra of

mn reactor fed with OA-SGW, plotted with uraninite

I water, plotted with standards for aqueous uranyl and

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 83994

sediments from the column reactor fed with SGW showed

similar results to those found in the column reactor fed with

DI water (Table 1, Fig. 2b and d).

The presence of cells embedded in EPS (i.e., sediment bio-

films) was observed in the sediment column reactor fed with

OA-SGW (Fig. S6). The organics supplied in OA-SGW stimu-

lated cell growth as well as the EPS production of indigenous

microorganisms in the 300 A sediments. The lactate concen-

tration in the effluent from the 300 A sediment column fed

with OA-SGW varied between 0 and 0.5 mM (data not shown);

therefore, w93% of the lactate delivered to the sediments was

consumed in the column. Furthermore, soluble iron and

manganese, presumably Fe(II) and Mn(II), were released into

the aqueous phase before and during U immobilization

(Fig. S7), indicating that anaerobic, metal-reducing conditions

had been established. Therefore, it is thought that the U

immobilization in this reactor resulted from a combination of

adsorption and reduction. Linear combination fitting of the U

LIII edge XANES data showed that 80e85% of the immobilized

U in both the inlet and outlet 300 A sediments from the

column reactor fed with OA-SGW was U(IV) (Table 1, Fig. 2a).

EXAFS analysis of the inlet and outlet 300 A sediments from

the column reactor fed with OA-SGW found nanoparticulate

uraninite (UO2) (Fig. 2c). The EXAFS data and fit of the inlet

300 A sediments from the column reactor fed with OA-SGW

are shown in Fig. S8. A comparison of the EXAFS analysis

(Table S2) of the current study with previously published

literature verifies the formation of nanoparticulate uraninite

with an average diameter of 1e2 nm (Boyanov et al., 2007). It is

interesting to note that the total amounts of U immobilized in

the 300 A sediments from the column reactors fed with SGW

and OA-SGW were comparable but that the U immobilization

in the column reactor fed with SGW was mostly through

adsorption as U(VI) rather than reduction to U(IV).

The U immobilization in the 300 A sediment column

reactor fed with OA-SGW was mainly through microbial

reduction to U(IV) (80e85%), suggesting that reductive immo-

bilization of U(VI) could be an important microbial process

affecting the fate and transport of U even under bulk oxic

conditions. The sediment sample was collected from a depth

of 33e34 ft. (w10 m), where bacterial richness was highest in

the Hanford formation (Lin et al., 2012a). Functional groups of

facultative and anaerobic respiring bacteria have been detec-

ted by molecular and cultivation approaches in samples from

the oxic Hanford formation, suggestingmetal reductionmight

occur in the oxic sediment (Lin et al., 2012b). Further, Lee et al.

(2012) have demonstrated the potential of microbial anaerobic

metabolism in shallow, saturated subsurface Hanford

formation sediments in the 300 A, suggesting that biogeo-

chemical hot spots might enhance microbial activity that

could result in local anoxia and reducing conditions (Lin et al.,

2012b; McClain et al., 2003; Nguyen et al., 2012). In such anoxic

microenvironment, certain indigenous microorganisms are

capable of immobilizing U(VI) through microbial enzymatic

reduction.We have isolated several bacterial strains including

Paenibacillus sp. 300 A from the sediments used in this study

and they are potentially involved in reductive immobilization

of U(VI), and further investigations are in progress. Microbial

reduction process could be affected by various factors during

the immobilization phase, including: (1) U toxicity inhibiting

cell growth. This could decrease the U reduction rate, extend

the reaction times, and even halt microbial growth and

metabolic activity (Sani et al., 2006). (2) The predominance of

calciumeuranylecarbonate complexes (w97%) significantly

decreasing the U reduction rate, since U(VI) is a much less

energetically favorable electron acceptor when it forms cal-

ciumeuranylecarbonate complexes (Brooks et al., 2003). (3)

The increased partial pressure of CO2 (pCO2) and increased

inorganic carbon ((bi)carbonate) concentrations resulting

from five months of stimulated microbial respiration prior to

U immobilization in 300 A sediments fed with OA-SGW (Wan

et al., 2005, 2008). Increased bicarbonate concentrations could

facilitate Fe(III)-driven U(IV) reoxidation and the formation of

higher-level uraniumecarbonate complexes while still under

reducing conditions (Ginder-Vogel et al., 2006; Wan et al.,

2005). (4) The presence of O2 in OA-SGW and of iron and

manganese in 300 A sediments, which could reoxidize the

reduced U(IV) abiotically. We observed soluble Fe and Mn

(quantified as total aqueous concentrations), presumably

Fe(II) and Mn(II) species, released from the reactor fed with

OA-SGW (Fig. S7). However, U remobilization by iron and

manganese requires physical contact between two solid

phases (e.g., nanoparticulate uraninite and solid Fe/Mn

minerals) (Fredrickson et al., 2002), which is unlikely to be the

case in unmixed sediment columns. Although all these

remobilization processes could occur, we observed contin-

uous U immobilization through microbial reduction when

300 A sediments were fed with OA-SGW.

3.2. U remobilization by oxic Columbia River water

Approximately 80e85% of U immobilization was due to

reduction in 300 A sediments fed with OA-SGW. If oxic

Columbia Riverwater causedU remobilization simply through

the desorption of U(VI), we would expect that 15e20% of the

adsorbed U associated with 300A sediments fed with OA-SGW

would be released into solutionwhen the samplewas exposed

to Columbia River water because of the changes in ground-

water composition and pH (seasonal variation w1.0 unit),

which can change U(VI) speciation (McKinley et al., 2011;

Poston et al., 2009; Yin et al., 2011). For example, Columbia

River water stage fluctuation (up to 1 m/d and >2 m per

season) alters groundwater flux on a diel basis and ground-

water chemistry on a seasonal basis at distances of 250 m and

more from the river (Hammond and Lichtner, 2010; Lindberg

and Petersen, 2004; Yin et al., 2011; Zachara et al., 2005).

Therefore, the variable groundwater compositions and pH and

the intrusion of Columbia River water into 300 A possibly

result in U remobilization through desorption. Therefore, the

majority of the U associated with 300 A sediments fed with

SGW or DI water was expected to be released when the

samples were exposed to Columbia River water, since U

immobilization was mainly through adsorption.

The percentage of U remobilized and the oxidation states

of the residual sediment-associated U are summarized in

Table 2. The introduction of Columbia River water resulted in

the release of U into solution from previously U-immobilized

300 A sediments fed with OA-SGW, SGW or DI water. The

percentage of U remaining in 300 A sediments after remobi-

lization using oxic Columbia River water is shown in Fig. 3. U

Table 2 e Percentage U remobilization during 25 days ofthe remobilization experiment and the oxidation statesof the residual sediment-associated U after 50 days of theremobilization experiment.

U remobilizationusingU-immobilized 300 Asediments fromcolumn reactorsfed with

% remobilization U oxidationstatea

OA-SGW w7% 100% U(IV)

SGW w7% 100% U(VI)

DI water w7% 100% U(VI)

a XANES analysis was carried out after 50 days of the remobiliza-

tion experiment (XANES availability). The sediments were fed with

Columbia River water throughout the time period.

Fig. 4 e U LIII edge XANES analysis of U-immobilized 300 A

sediments fed with OA-SGW after remobilization by oxic

Columbia River water.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 8 3995

remobilization from 300 A sediments fed with OA-SGW

stopped within eight days and approximately 93% of the U

remained associated with the sediments, whereas linear

remobilization trends were observed for 300 A sediments fed

with SGW or DI water. We have shown that the U immobili-

zation in 300 A sediments fed with SGW or DI water was

mostly due to adsorption (Fig. 2b and d); hence, the remobili-

zation of U from these 300 A sediments can be mainly attrib-

uted to the desorption of U(VI). The concentrations of Ca2þ

(w17 mg/L) and CO2�3 (w68 mg/L) in Columbia River water at

pH (w7.8) could lead to desorption of the U from U-immobi-

lized 300 A sediments (Table S3) (Poston et al., 2009; Yin et al.,

2011). In contrast, the U immobilization in 300 A sediments fed

with OA-SGWwas due to both U(VI) adsorption and reduction

to U(IV) (Fig. 2a); therefore, the remobilization of approxi-

mately 7% of the total immobilized U in 300 A sediments fed

with OA-SGW could be attributed to reoxidation, desorption

and/or detachment of the U-immobilized biomass, since we

measured the total U concentration in unfiltered column

reactor effluent. U LIII edge XANES analysis of U associated

with 300 A sediments fed with OA-SGW after exposure to

Columbia River water (Fig. 4) showed that most of the

Fig. 3 e Percentage U remaining in 300 A sediments after

remobilization using oxic Columbia River water. The 300 A

sediments used for remobilization were from U-

immobilized 300 A sediments fed with OA-SGW (>), SGW

(D) or DI water (,) in column reactors.

remaining U (w93%) remained in the form of U(IV) nano-

particles, suggesting that most of the U immobilized in these

sediments was recalcitrant to reoxidation. The Fourier trans-

form of U LIII-edge EXAFS data with biogenic nanoparticulate

uraninite and monomeric U(IV) standards confirmed that the

U in the inlet and outlet samples fed with Columbia River

water was in the form of nanoparticulate uraninite (Fig. S9).

At the end of the remobilization experiment (w50 days)

with Columbia River water, DO concentrations in the outlet of

the column containing U immobilized during the OA-SGW-

stimulated phase were 5.21e5.91 mg/L. Approximately

25e35% of the DO in the oxic Columbia River water was

consumed by the reduced sediments through biotic and/or

abiotic processes.Although the oxidationofU(IV) bymolecular

O2 is thermodynamically favorable (Abdelouas et al., 1999) and

hasbeendocumentedasamechanismforU(IV) remobilization

in previous lab-scale and field studies (Casas et al., 1994; Moon

et al., 2007; Zhong et al., 2005), we found that U remained as

U(IV) nanoparticles in 300A sedimentsmicrobially reduced via

stimulation with OA-SGW and subsequently exposed to oxic

Columbia River water (w80 pore volumes). This is consistent

with our hypothesis that, under bulk aerobic conditions,

microbial growth in the 300A subsurface sediments can create

anoxic zones and prevent the reoxidation of reduced U(IV)

species (Krawczyk-Barsch et al., 2008).We should note that we

have recently found anoxic microenvironments inside sedi-

ment biofilms from 300 A, suggesting that U(VI) reduction

proceeds under bulk oxic conditions (Nguyen et al., 2012).

Moreover, we should note that river water may also introduce

organics (Becker and Gray, 1992; Bisping, 2010; Moser et al.,

2003) to stimulate and maintain microbial activity which

keeps uranium in a reduced state.

4. Conclusions

Our study demonstrates U immobilization and remobilization

in Hanford 300 A sediments in the presence of oxic

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 83996

groundwater and Columbia River water, mimicking the

interactions among U, groundwater or Columbia River water,

and sediments in the subsurface environment of Hanford

300 A. Our results suggest that U can be immobilized by OA-

SGW stimulation of indigenous microorganisms in sedi-

ments in the Hanford 300 A subsurface, predominantly

through reduction, even in the presence of oxic groundwater.

The reduced U in the 300 A sediments fed with OA-SGW was

resistant to reoxidation by oxic Columbia River water.

Columbia River water caused U remobilization (w7%) through

desorption, and most of the U that remained in the 300 A

sediments (w93%) fed with OA-SGW was in the form of

uraninite nanoparticles. Taken together, these findings imply

that: 1) the reductive immobilization of U through OA-SGW

stimulation of indigenous 300 A sediment microorganisms

could be viable in the relatively oxic Hanford 300 A subsurface

environments; and 2) with the intrusion of oxic Columbia

River water, desorption might be an important process

contributing to U remobilization from OA-SGW-stimulated

sediments in the Hanford 300 A subsurface.

Acknowledgments

This researchwas supported by the U.S. Department of Energy

(DOE) Office of Biological and Environmental Research (BER)

under the Subsurface Biogeochemical Research (SBR) Program

(grant DE-FG92-08ER64560). PNNL contributions to this

research were supported in part by the PNNL Scientific Focus

Area and Hanford 300A IFRC projects, and ANL contributions

were supported in part by the ANL Scientific Focus Area

project, which are part of the SBR Program of the Office of

Biological and Environmental Research (BER), U.S. DOE under

contract DE-AC05-76RLO and DE-AC02-06CH11357, respec-

tively. Use of the Advanced Photon Source (APS) was sup-

ported by the DOE-SC Office of Basic Energy Sciences, under

contract DE-AC02-06CH11357. MRCAT/EnviroCAT operations

are supported by DOE and the MRCAT/EnviroCAT member

institutions. The U.S. government retains for itself and others

acting on its behalf a paid-up, non-exclusive, irrevocable

worldwide license in said article to reproduce, prepare deriv-

ative works, distribute copies to the public and perform

publicly and display publicly, by or on behalf of the Govern-

ment. We are also grateful to the Franceschi Microscopy and

Imaging Center of Washington State University for the use of

their facilities and staff assistance.

Appendix A. Supplementary material

Supplementary material associated with this article can be

found, in the online version, at doi:10.1016/j.watres.2012.05.027.

r e f e r e n c e s

Abdelouas, A., Lu, Y.M., Lutze, W., Nuttall, H.E., 1998. Reductionof U(VI) to U(IV) by indigenous bacteria in contaminated

ground water. Journal of Contaminant Hydrology 35 (1e3),217e233.

Abdelouas, A., Lutze, W., Nuttall, H.E., 1999. Uraniumcontamination in the subsurface: characterization andremediation. Reviews in Mineralogy 38, 433e473.

Anderson, R.T., Vrionis, H.A., Ortiz-Bernad, I., Resch, C.T.,Long, P.E., Dayvault, R., Karp, K., Marutzky, S., Metzler, D.R.,Peacock, A., White, D.C., Lowe, M., Lovley, D.R., 2003.Stimulating the in situ activity of Geobacter species to removeuranium from the groundwater of a uranium-contaminatedaquifer. Applied and Environmental Microbiology 69 (10),5884e5891.

Ankudinov, A., Ravel, B., Rehr, J., Conradson, S., 1998. Real-spacemultiple-scattering calculation and interpretation of x-ray-adsorption near-edge structure. Physical Review B 58,7565e7576.

Becker, C.D., Gray, R.H., 1992. Past and present water-qualityconditions in the Hanford Reach, Columbia River.Environmental Monitoring and Assessment 22 (2), 137e152.

Bernier-Latmani, R., Veeramani, H., Vecchia, E.D., Junier, P.,Lezama-Pacheco, J.S., Suvorova, E.I., Sharp, J.O.,Wigginton, N.S., Bargar, J.R., 2010. Non-uraninite products ofmicrobial U(VI) reduction. Environmental Science &Technology 44 (24), 9456e9462.

Bisping, L.E., 2010. Hanford Site Environmental SurveillanceData Report for Calendar Year 2009. PNNL-19455, APP.1,Richland, WA.

Bjornstad, B.N., Horner, J.A., 2008. Drilling, Sampling, and Well-Installation Plan for the IFC Well Field, 300 Area. PNNL-17512,Richland, WA.

Bjornstad, B.N., Lanigan, D.C., Horner, J.A., Thorne, P.D.,Vermeul, V.R., 2009. Borehole Completion and ConceptualHydrogeologic Model for the IFRC Well Field, 300 Area.Hanford site, Richland, WA. PNNL-18340.

Boyanov, M.I., Fletcher, K.E., Kwon, M.J., Rui, X., O’Loughlin, E.J.,Loffler, F.E., Kemner, K.M., 2011. Solution and microbialcontrols on the formation of reduced U(IV) species.Environmental Science & Technology 45 (19), 8336e8344.

Boyanov, M.I., O’Loughlin, E.J., Roden, E.E., Fein, J.B.,Kemner, K.M., 2007. Adsorption of Fe(II) and U(VI) to carboxyl-functionalized microspheres: the influence of speciation onuranyl reduction studied by titration and XAFS. Geochimica etCosmochimica Acta 71 (8), 1898e1912.

Brooks, S.C., Fredrickson, J.K., Carroll, S.L., Kennedy, D.W.,Zachara, J.M., Plymale, A.E., Kelly, S.D., Kemner, K.M.,Fendorf, S., 2003. Inhibition of bacterial U(VI) reduction bycalcium. Environmental Science & Technology 37 (9),1850e1858.

Brown, C.F., Um, W., Serne, R.J., 2008. Uranium Contamination inthe 300 Area: Emergent Data and Their Impact on the SourceTerm Conceptual Model. PNNL-17793, Richland, WA.

Burgos, W.D., McDonough, J.T., Senko, J.M., Zhang, G.,Dohnalkova, A.C., Kelly, S.D., Gorby, Y., Kemner, K.M., 2008.Characterization of uraninite nanoparticles produced byShewanella oneidensis MR-1. Geochimica et Cosmochimica Acta72 (20), 4901e4915.

Cao, B., Ahmed, B., Kennedy, D.W., Wang, Z.M., Shi, L.,Marshall, M.J., Fredrickson, J.K., Isern, N.G., Majors, P.D.,Beyenal, H., 2011a. Contribution of extracellular polymericsubstances from Shewanella sp. HRCR-1 biofilms to U(VI)immobilization. Environmental Science & Technology 45 (13),5483e5490.

Cao, B., Shi, L.A., Brown, R.N., Xiong, Y.J., Fredrickson, J.K.,Romine, M.F., Marshall, M.J., Lipton, M.S., Beyenal, H., 2011b.Extracellular polymeric substances from Shewanella sp.HRCR-1 biofilms: characterization by infrared spectroscopyand proteomics. Environmental Microbiology 13 (4),1018e1031.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 8 3997

Casas, I., Gimenez, J., Marti, V., Torrero, M.E., Depablo, J., 1994.Kinetics studies of unirradiated UO2 dissolution underoxidizing conditions in batch and flow experiments.Radiochimica Acta 66-7, 23e27.

Dalla Vecchia, E.C., Veeramani, H., Suvorova, E.I., Wigginton, N.S.,Bargar, J.R., Bernier-Latmani, R., 2010. U(VI) reduction byspores of Clostridium acetobutylicum. Research in Microbiology161 (9), 765e771.

Decho, A.W., 2000. Microbial biofilms in intertidal systems: anoverview. Continental Shelf Research 20 (10e11), 1257e1273.

Fletcher, K.E., Boyanov, M.I., Thomas, S.H., Wu, Q., Kemner, K.M.,Loeffler, F.E., 2010. U(VI) reduction to mononuclear U(IV) byDesulfitobacterium species. Environmental Science &Technology 44 (12), 4705e4709.

Fox, P.M., Davis, J.A., Zachara, J.M., 2006. The effect of calcium onaqueous uranium(VI) speciation and adsorption to ferrihydriteand quartz. Geochimica et Cosmochimica Acta 70 (6),1379e1387.

Fredrickson, J.K., Zachara, J.M., Kennedy, D.W., Liu, C.X.,Duff, M.C., Hunter, D.B., Dohnalkova, A., 2002. Influence of Mnoxides on the reduction of uranium(VI) by the metal-reducingbacterium Shewanella putrefaciens. Geochimica etCosmochimica Acta 66 (18), 3247e3262.

Ginder-Vogel, M., Criddle, C.S., Fendorf, S., 2006. Thermodynamicconstraints on the oxidation of biogenic UO2 by Fe(III) (Hydr)oxides. Environmental Science & Technology 40 (11),3544e3550.

Gorman-Lewis, D., Elias, P.E., Fein, J.B., 2005. Adsorption ofaqueous uranyl complexes onto Bacillus subtilis cells.Environmental Science & Technology 39 (13), 4906e4912.

Hammond, G.E., Lichtner, P.C., 2010. Field-scale model for thenatural attenuation of uranium at the Hanford 300 Area usinghigh-performance computing. Water Resources Research 46.

Hedaya, M.A., Birkenfeld, H.P., Kathren, R.L., 1997. A sensitivemethod for the determination of uranium in biologicalsamples utilizing kinetic phosphorescence analysis (KPA).Journal of Pharmaceutical and Biomedical Analysis 15 (8),1157e1165.

Istok, J.D., Senko, J.M., Krumholz, L.R., Watson, D., Bogle, M.A.,Peacock, A., Chang, Y.J., White, D.C., 2004. In situ bioreductionof technetium and uranium in a nitrate-contaminated aquifer.Environmental Science & Technology 38 (2), 468e475.

Jones, D.S., Tobler, D.J., Schaperdoth, I., Mainiero, M.,Macalady, J.L., 2010. Community structure of subsurfacebiofilms in the thermal sulfidic caves of Acquasanta Terme,Italy. Applied and Environmental Microbiology 76 (17),5902e5910.

Kemner, K.M., Kelly, S.D., 2007. In: Hurst, C.J. (Ed.), Manual ofEnvironmental Microbiology. ASM Press, pp. 1183e1194.

Krawczyk-Barsch, E., Grossmann, K., Arnold, T., Hofmann, S.,Wobus, A., 2008. Influence of uranium (VI) on the metabolicactivity of stable multispecies biofilms studied by oxygenmicrosensors and fluorescence microscopy. Geochimica etCosmochimica Acta 72 (21), 5251e5265.

Lee, J.-H., Fredrickson, J.K., Kukkadapu, R.K., Boyanov, M.I.,Kemner, K.M., Lin, X., Kennedy, D.W., Bjornstad, B.N.,Konopka, A.E., Moore, D.A., Resch, C.T., Phillips, J.L., 2012.Microbial reductive transformation of Phyllosilicate Fe(III)and U(VI) in fluvial subsurface sediments. EnvironmentalScience & Technology 46 (7), 3721e3730.

Lewandowski, Z., Beyenal, H., 2007. Foundamentals of BiofilmsResearch. CRC Press.

Lin, X., Kennedy, D., Fredrickson, J., Bjornstad, B., Konopka, A.,2012a. Vertical stratification of subsurface microbialcommunity composition across geological formations at theHanford Site. Environmental Microbiology 14 (2), 414e425.

Lin, X., Kennedy, D., Peacock, A., McKinley, J., Resch, C.T.,Fredrickson, J., Konopka, A., 2012b. Distribution of microbial

biomass and potential for anaerobic respiration in Hanfordsite 300 area subsurface sediment. Applied and EnvironmentalMicrobiology 78 (3), 759e767.

Lindberg, M.J., Petersen, R.E., 2004. 300-FF-5 Operable Unit,Hanford Site Groundwater Monitoring for Fiscal Year 2004.PNNL-15070.

Lovley, D.R., 1993. Dissimilatory metal reduction. Annual Reviewof Microbiology 47, 263e290.

Lovley, D.R., Phillips, E.J.P., Gorby, Y.A., Landa, E.R., 1991.Microbial reduction of uranium. Nature 350 (6317),413e416.

Marshall, M.J., Beliaev, A.S., Dohnalkova, A.C., Kennedy, D.W.,Shi, L., Wang, Z.M., Boyanov, M.I., Lai, B., Kemner, K.M.,McLean, J.S., Reed, S.B., Culley, D.E., Bailey, V.L.,Simonson, C.J., Saffarini, D.A., Romine, M.F., Zachara, J.M.,Fredrickson, J.K., 2006. c-Type cytochrome-dependentformation of U(IV) nanoparticles by Shewanella oneidensis. PlosBiology 4 (8), 1324e1333.

Marshall, M.J., Dohnalkova, A.C., Kennedy, D.W., Plymale, A.E.,Thomas, S.H., Loffler, F.E., Sanford, R.A., Zachara, J.M.,Fredrickson, J.K., Beliaev, A.S., 2009. Electron donor-dependentradionuclide reduction and nanoparticle formation byAnaeromyxobacter dehalogenans strain 2CP-C. EnvironmentalMicrobiology 11 (2), 534e543.

Marsili, E., Beyenal, H., Di Palma, L., Merli, C., Dohnalkova, A.,Amonette, J.E., Lewandowski, Z., 2005. Uranium removal bysulfate reducing biofilms in the presence of carbonates. WaterScience and Technology 52 (7), 49e55.

Marsili, E., Beyenal, H., Palma, L.D., Merli, C., Dohnalkova, A.,Amonette, J.E., Lewandowski, Z., 2007. Uraniumimmobilization by sulfate-reducing biofilms grown onHematite, Dolomite, and Calcite. Environmental Science &Technology 41 (24), 8349e8354.

McClain, M.E., Boyer, E.W., Dent, C.L., Gergel, S.E., Grimm, N.B.,Groffman, P.M., Hart, S.C., Harvey, J.W., Johnston, C.A.,Mayorga, E., McDowell, W.H., Pinay, G., 2003. Biogeochemicalhot spots and hot moments at the interface of terrestrial andaquatic ecosystems. Ecosystems 6 (4), 301e312.

McKinley, J.P., Zachara, J.M., Resch, C.T., Kaluzny, R.M.,Miller, M.D., Vermeul, V.R., Fritz, B.G., Moser, J.V., 2011. RiverWater Intrusion and Contaminant Uranium Contributionsfrom the Vadose Zone to Groundwater during the AnnualSpring Rise in Columbia River Stage at the Hanford Site 300Area. Washington. Environmental Science & TechnologySubmitted.

McKinley, J.P., Zachara, J.M., Wan, J., McCready, D.E., Heald, S.M.,2007. Geochemical controls on contaminant uranium invadose Hanford formation sediments at the 200 area and 300area, Hanford Site, Washington. Vadose Zone Journal 6 (4),1004e1017.

McLean, J.S., Majors, P.D., Reardon, C.L., Bilskis, C.L., Reed, S.B.,Romine, M.F., Fredrickson, J.K., 2008. Investigations ofstructure and metabolism within Shewanella oneidensis MR-1biofilms. Journal of Microbiological Methods 74 (1), 47e56.

Moon, H.S., Komlos, J., Jaffe, P.R., 2009. Biogenic U(IV) oxidation bydissolved oxygen and nitrate in sediment after prolongedU(VI)/Fe(III)/SO42- reduction. Journal of ContaminantHydrology 105 (1e2), 18e27.

Moon, H.S., Komlos, J., Jaffe, P.R., 2007. Uranium reoxidation inpreviously bioreduced sediment by dissolved oxygen andnitrate. Environmental Science & Technology 41 (13),4587e4592.

Moser, D.P., Fredrickson, J.K., Geist, D.R., Arntzen, E.V.,Peacock, A.D., Li, S.M.W., Spadoni, T., McKinley, J.P., 2003.Biogeochemical processes and microbial characteristicsacross groundwater-surface water boundaries of the HanfordReach of the Columbia River. Environmental Science &Technology 37 (22), 5127e5134.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 3 9 8 9e3 9 9 83998

N’Guessan, A.L., Vrionis, H.A., Resch, C.T., Long, P.E., Lovley, D.R.,2008. Sustained removal of uranium from contaminatedgroundwater following stimulation of dissimilatory metalreduction. Environmental Science & Technology 42 (8),2999e3004.

Nguyen, H.D., Cao, B., Mishra, B., Boyanov, M.I.,Kemner, K.M., Fredrickson, J.K., Beyenal, H., 2012.Microscale geochemical gradients in Hanford 300 Areasediment biofilms and influence of uranium. WaterResearch 46 (1), 227e234.

North, N.N., Dollhopf, S.L., Petrie, L., Istok, J.D., Balkwill, D.L.,Kostka, J.E., 2004. Change in bacterial community structureduring in situ biostimulation of subsurface sedimentcocontaminated with uranium and nitrate. Applied andEnvironmental Microbiology 70 (8), 4911e4920.

Palmisano, A., Hazen, T., 2003. Bioremediation of Metals andRadionuclides: What It Is and How It Works, second ed..LBNL-42595, Berkeley, CA.

Peterson, R.E., 2001. Zone of Interaction between Hanford SiteGroundwater and Adjacent Columbia River. PNNL-13674,Richland, WA.

Peterson, R.E., Rockhold, M.L., Serne, R.J., Thorne, P.D.,Williams, M.D., 2008. Uranium Contamination in theSubsurface beneath the 300 Area. Hanford site, Richland, WA.Washington, PNNL-17034.

Poston, T.M., Duncan, J.P., Dirkes, R.L., 2009. Hanford SiteEnvironmental Report for Calendar Year 2008. PNNL-18427,Richland, WA.

Ravel, B., Newville, M., 2005. ATHENA, ARTEMIS, HEPHAESTUS:data analysis for x-ray absorption spectroscopy using IFEFFIT.Journal of Synchrotron Radiation 12, 537e541.

Redden, G., Bargar, J., Bencheikh-Latmani, R., 2001. Citrateenhanced uranyl adsorption on goethite: an EXAFSanalysis. Journal of Colloid and Interface Science 244 (1),211e219.

Riley, R.G., Zachara, J.M., 1992. Chemical Contaminants onDOE Lands and Selection of Contaminant Mixturesfor Subsurface Science Research. DOE/ER-0547T,Richland, WA.

Sani, R.K., Peyton, B.M., Dohnalkova, A., 2006. Toxic effects ofUranium on Desulfovibrio desulfuricans G20. EnvironmentalToxicology and Chemistry 25 (5), 1231e1238.

Sani, R.K., Peyton, B.M., Dohnalkova, A., Amonette, J.E., 2005.Reoxidation of reduced uranium with iron(III) (hydr)oxidesunder sulfate-reducing conditions. Environmental Science &Technology 39 (7), 2059e2066.

Sani, R.K., Peyton, B.M., Smith, W.A., Apel, W.A., Petersen, J.N.,2002. Dissimilatory reduction of Cr(VI), Fe(III), and U(VI) byCellulomonas isolates. Applied Microbiology and Biotechnology60 (1e2), 192e199.

Segre, C., Leyarovska, N., Chapman, L., Lavender, W., Plag, P.,King, A., Kropf, A., Bunker, B., Kemner, K., Dutta, P., Duran, R.,Kaduk, J., 2000. The MRCAT Inserton Device Beamline at theAdvanced Photon Source, pp. 419e422.

Senko, J.M., Istok, J.D., Suflita, J.M., Krumholz, L.R., 2002.In-situ evidence for uranium immobilization andremobilization. Environmental Science & Technology 36 (7),1491e1496.

Sivaswamy, V., Boyanov, M.I., Peyton, B.M., Viamajala, S.,Gerlach, R., Apel, W.A., Sani, R.K., Dohnalkova, A.,Kemner, K.M., Borch, T., 2011. Multiple mechanisms ofUranium immobilization by Cellulomonas sp. strain ES6.Biotechnology and Bioengineering 108 (2), 264e276.

Stern, E., Newville, M., Ravel, B., Yacoby, Y., Haskel, D., 1995. TheUWXAFS analysis package- philosophy and details. Physica B209, 117e120.

Stewart, B.D., Mayes, M.A., Fendorf, S., 2010. Impact ofuranylecalciumecarbonato complexes on uranium(VI)adsorption to synthetic and natural sediments. EnvironmentalScience & Technology 44 (3), 928e934.

Stoliker, D.L., Kent, D.B., Zachara, J.M., 2011. Quantifyingdifferences in the impact of variable chemistry on equilibriumuranium(VI) adsorption properties of aquifer sediments.Environmental Science & Technology 45 (20), 8733e8740.

Suzuki, Y., Kelly, S.D., Kemner, K.M., Banfield, J.F., 2002.Radionuclide contamination e Nanometre-size products ofUranium bioreduction. Nature 419 (6903), 134.

Tokunaga, T.K., Wan, J., Kim, Y., Sutton, S.R., Newville, M.,Lanzirotti, A., Rao, W., 2008. Real-time X-ray absorptionspectroscopy of uranium, iron, and manganese incontaminated sediments during bioreduction. EnvironmentalScience & Technology 42 (8), 2839e2844.

Wan, J., Tokunaga, T.K., Brodie, E., Wang, Z., Zheng, Z.,Herman, D., Hazen, T.C., Firestone, M.K., Sutton, S.R., 2005.Reoxidation of bioreduced uranium under reducingconditions. Environmental Science & Technology 39 (16),6162e6169.

Wan, J., Tokunaga, T.K., Kim, Y., Brodie, E., Daly, R., Hazen, T.C.,Firestone, M.K., 2008. Effects of organic carbon supply rates onuranium mobility in a previously bioreduced contaminatedsediment. Environmental Science & Technology 42 (20),7573e7579.

Wazne, M., Korfiatis, G.P., Meng, X., 2003. Carbonate effects onhexavalent uranium adsorption by iron oxyhydroxide.Environmental Science & Technology 37 (16), 3619e3624.

Weidler, G.W., Dornmayr-Pfaffenhuemer, M., Gerbl, F.W.,Heinen, W., Stan-Lotter, H., 2007. Communities of archaea andbacteria in a subsurface radioactive thermal spring in theAustrian Central Alps, and evidence of ammonia-oxidizingCrenarchaeota. Applied and Environmental Microbiology 73 (1),259e270.

Wyckoff, R., 1960. Crystal Structures. Interscience, New York.Yabusaki, S.B., Fang, Y.L., Waichler, S.R., 2008. Building

conceptual models of field-scale uranium reactive transport ina dynamic Vadose zone-aquifer-river system. WaterResources Research 44, W12403.

Yin, J.Y.J., Haggerty, R., Stoliker, D.L., Kent, D.B., Istok, J.D.,Greskowiak, J., Zachara, J.M., 2011. Transient groundwaterchemistry near a river: effects on U(VI) transport inlaboratory column experiments. Water Resources Research47, W04502.

Zachara, J.M., 2009. Multi-scale Mass Transfer ProcessesControlling Natural Attenuation and Engineered Remediation:An IFRC Focused on Hanford’s 300 Area Uranium Plume.PNNL-SA-64785, Richland, WA.

Zachara, J.M., Davis, J.A., Liu, C., McKinley, J.P., Qafoku, N.,Wellman, D.M., Yabusaki, S.B., 2005. Uranium Geochemistryin Vadose Zone and Aquifer Sediments from the 300 AreaUranium Plume. PNNL-15121, Richland, WA.

Zheng, Z., Tokunaga, T.K., Wan, J., 2003. Influence of calciumcarbonate on U(VI) sorption to soils. Environmental Science &Technology 37 (24), 5603e5608.

Zhong, L.R., Liu, C.X., Zachara, J.M., Kennedy, D.W., Szecsody, J.E.,Wood, B., 2005. Oxidative remobilization of biogenicuranium(IV) precipitates: effects of iron(II) and pH. Journal ofEnvironmental Quality 34 (5), 1763e1771.