Assessing the Potential for 

Biorestoration of Uranium 

In Situ Recovery Sites

Office of Nuclear Regulatory Research

NUREG/CR-7167
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Assessing the Potential for 

Biorestoration of Uranium 

In Situ Recovery Sites

Manuscript Completed: September 2013

Date Published: June 2014

Prepared by:

S. B. Yabusaki1, Y. Fang1, S. R. Waichler1, C. C. Fuller2,

K. Akstin2

, P. E. Long3, and M. Fuhrmann4

1

Pacific Northwest National Laboratory

2

U.S. Geological Survey

3

Lawrence Berkeley National Laboratory

4

U.S. Nuclear Regulatory Commission

Mark Fuhrmann, NRC Project Manager

Office of Nuclear Regulatory Research

NUREG/CR-7167

iii

Abstract

In situ bioremediation (ISB) may be a more effective alternative for the restoration of

uranium in situ recovery (ISR) sites than current methods. ISB involves the stimulation of 

indigenous microorganisms that catalyze the immobilization of targeted contaminants. Many of 

the post-operational contaminants found at ISR sites (e.g., uranium, vanadium, selenium, 

molybdenum) are in oxidized, soluble forms. This is a result of the ISR process in which 

oxidizing solutions are injected into the subsurface to liberate uranium from its solid matrix. In 

this case, ISB would be directed at the chemical reduction of these contaminants to their 

immobile forms. While the technique has the potential for more effective restoration, shorter 

restoration periods, and lower cost, it is an unproven technology that relies on the stability of 

uranium and other metals that are left in place. Consequently, there is a need to better 

understand the behavior of the uranium in the bioremediated system during and after the 

restoration. Licensees have been conducting studies into the effectiveness of ISB, therefore an 

independent examination, such as that documented in this report, is important for anticipated 

regulatory decisionmaking regarding its acceptability. 

This report is the third in a series of technical NUREG/CRs that are intended to provide 

background and evaluation of uranium ISB technology. The purpose of this report is to assess 

the potential effectiveness of ISB for uranium ISR sites, identify performance issues, and guide 

future implementation and monitoring. To address the limitations of the current knowledge 

base of field-scale ISB uranium behavior, the approach in this document is to augment the 

available geological, hydrological, and chemical data from existing ISR sites with 1) laboratory 

studies of biologically mediated removal of uranium from solution using sediment taken from a 

leached out uranium ore body, 2) previous studies of uranium ISB in shallow, alluvial aquifers, 

and 3) coupled process modeling of hypothetical field-scale ISR restoration including: coupled 

flow, transport, and biogeochemical processes. The models are intended to help integrate and 

scale up the predicted effects from the laboratory to the field. The experimental and modeling 

studies presented here suggest that biostimulation of indigenous bacterial populations can be 

effective in lowering aqueous concentrations of uranium at ISR sites to acceptable levels.

v 

Contents

Abstract .......................................................................................................................... iii

Contents .......................................................................................................................... v

Figures............................................................................................................................ ix

Tables .......................................................................................................................... xvii

Executive Summary..................................................................................................... xix

Acknowledgments ....................................................................................................... xxi

Acronyms and Abbreviations ....................................................................................xxiii

1 Introduction .......................................................................................................... 1-1

1-1 Demand for Uranium....................................................................................... 1-1

1.1 Uranium Mining in the U.S............................................................................... 1-2

1.1.1 Uranium Ore Deposits ........................................................................... 1-2

1.2 In Situ Recovery of Uranium............................................................................ 1-3

1.2.1 Requirements for ISR ............................................................................ 1-5

1.2.2 Lixiviants for ISR.................................................................................... 1-5

1.2.3 ISR Life Cycle........................................................................................ 1-6

1.2.4 ISR Impacts and Containment Issues.................................................... 1-7

1.3 Restoration of Uranium ISR Sites.................................................................... 1-7

1.3.1 General Description............................................................................... 1-7

1.3.2 Restoration Processes and Phases....................................................... 1-9

1.3.3 Regulatory Standards for Restoration.................................................. 1-11

1.4 Bioremediation .............................................................................................. 1-12

1.5 Document Description................................................................................... 1-14

1.5.1 Document Purpose, Approach, and Objectives ................................... 1-14

1.5.2 Document Contents............................................................................. 1-15

2 Geochemistry and Restoration Issues................................................................ 2-1

2.1 Geochemistry of Roll Front Deposits ............................................................... 2-1

2.2 Geochemistry of Uranium ISR......................................................................... 2-2

2.3 ISR Restoration Issues.................................................................................... 2-2

2.4 Uranium Bioremediation.................................................................................. 2-4

2.5 Findings from Uranium Bioremediation Field Research in Shallow Aquifers.... 2-6

2.6 Insights from Field Studies .............................................................................. 2-9

2.7 ISR Biorestoration Potential .......................................................................... 2-10

3 Laboratory Studies of Biostimulated Uranium Reduction in ISR Aquifer 

Sediments and Suboxic Remobilization of Sequestered Uranium 

(C.C. Fuller and K. Akstin).................................................................................... 3-1

3.1 Introduction ..................................................................................................... 3-1
vi

3.1.1 Objectives.............................................................................................. 3-1

3.1.2 Kingsville ISR Site Description............................................................... 3-1

3.2 Methods .......................................................................................................... 3-6

3.2.1 Sediment and Groundwater Collection and Analyses ............................ 3-6

3.2.2 AGW Design and Composition .............................................................. 3-9

3.2.3 Batch Experiments to Determine Electron Donors............................... 3-10

3.2.4 Column Set Up and Operation............................................................. 3-12

3.2.5 Column Effluent................................................................................... 3-15

3.2.6 Solid Phase Sampling ......................................................................... 3-18

3.2.7 Microbial Assay ................................................................................... 3-21

3.2.8 Spectroscopic Measurements.............................................................. 3-23

3.3 Results.......................................................................................................... 3-26

3.3.1 Batch Experiments .............................................................................. 3-26

3.3.2 Column Effluent................................................................................... 3-31

3.3.3 Solid Phase Characterization of Column Sediments............................ 3-43

3.3.4 X-ray Spectroscopy and μXRF Mapping of Column Sediment............. 3-54

3.3.5 Microbial Assay ................................................................................... 3-78

3.4 Discussion of Experimental Findings............................................................. 3-84

3.4.1 Biogeochemical Processes During Biostimulation ............................... 3-84

3.4.2 Stability of Immobilized Uranium.......................................................... 3-90

3.4.3 Comparison to Model Simulation of Biostimulated Reduction .............. 3-93

3.5 Summary of Experimental Findings............................................................... 3-94

4 Simulation of Uranium Bioremediation in ISR Sediment Columns................... 4-1

4.1 Uranium Bioremediation Modeling Studies...................................................... 4-1

4.2 Findings from Column Experiments using Kingsville Dome ISR Sediments .... 4-2

4.3 Conceptual Biorestoration Model..................................................................... 4-7

4.4 Simulator Description ...................................................................................... 4-9

4.5 Simulation Description..................................................................................... 4-9

4.6 Biologically-Mediated Processes: Shallow Groundwater Plume versus 

Deep ISR ...................................................................................................... 4-12

4.7 Model Implementation................................................................................... 4-13

4.7.1 Uranium Speciation ............................................................................. 4-15

4.7.2 Major Ion Reactions............................................................................. 4-16

4.7.4 Kinetic Reactions................................................................................. 4-17

4.8 Biostimulation Modeling Approach for the Column Experiments.................... 4-18

4.9 Column Simulation Results ........................................................................... 4-18

4.10Biostimulation Studies: Deep ISR vs Shallow Aquifer................................... 4-27

5 Field-Scale Flow and Transport Modeling .......................................................... 5-1

5.1 Overview......................................................................................................... 5-1
vii

5.2 Field-Scale Coupled Process Simulator .......................................................... 5-2

5.2.1 Physical System .................................................................................... 5-4

5.2.2 Model Flow System ............................................................................... 5-5

5.2.3 Flow and Transport Results................................................................. 5-10

5.3 Findings ........................................................................................................ 5-23

6 Field-Scale Post-ISR Biorestoration Modeling ................................................... 6-1

6.1 Biorestoration Concepts .................................................................................. 6-1

6.2 Modeling Approach ......................................................................................... 6-2

6.3 Field-Scale Biorestoration Modeling Results ................................................... 6-3

6.3.1 10 Days ................................................................................................. 6-8

6.3.2 30 Days ............................................................................................... 6-16

6.3.3 Modeling Results: Injectate without Terminal Electron Acceptors ........ 6-26

6.4 Background Injectate..................................................................................... 6-27

6.4.1 Lactate ................................................................................................ 6-27

6.4.2 Acetate................................................................................................ 6-28

6.4.3 U(VI).................................................................................................... 6-28

6.4.4 Adsorbed U(VI).................................................................................... 6-29

6.4.5 U(IV).................................................................................................... 6-30

6.4.6 Fe(II).................................................................................................... 6-30

6.4.7 Phyllosilicate Fe(III) ............................................................................. 6-31

6.4.8 Goethite............................................................................................... 6-31

6.4.9 FeS...................................................................................................... 6-32

6.4.10 Sulfate ................................................................................................ 6-32

6.4.11 Sulfide ................................................................................................ 6-33

6.5 Summary of Field-Scale Model Findings ....................................................... 6-34

7 Conclusions.......................................................................................................... 7-1

7.1 Considerations ................................................................................................ 7-4

7.2 Potential Approach.......................................................................................... 7-4

7.2.1 Field Characterization and Monitoring ................................................... 7-5

7.2.3 Long-Term Monitoring ......................................................................... 7-11

8 Reference List....................................................................................................... 8-1

ix

Figures 

Figure 1-1 Weighted-average price of uranium purchased by owners and operators of 

U.S. civilian nuclear power reactors, 1994-2011 deliveries. ................................... 1-1

Figure 1-2 Uranium roll front development in confined sandstone interval showing 

transport of oxidized U and its’ precipitation in the reduced zone (NRC, 

1997)........................................................................................................................ 1-3

Figure 1-3 Schematic diagram of typical 5-spot and 7-spot injection/production well 

patterns (NRC, 1997). .............................................................................................. 1-4

Figure 1-4 Conceptual representation of the ISR process, operations, and facilities 

(illustration reproduced with permission from the World Nuclear 

Association). ............................................................................................................ 1-6

Figure 2-1 Downgradient metal concentrations responding to acetate injection at the 

Rifle IFRC site. Concentrations were measured in well CD-01 from June 2010 

to May 2012. (Courtesy of Ken Williams, LBNL)................................................... 2-7

Figure 3-1 Location of Kingsville Dome ISR site in south Texas................................................ 3-3

Figure 3-2 Samples were obtained by drilling into an ore zone after uranium recovery 

operations were finished......................................................................................... 3-4

Figure 3-3 The sediment sample used for the bioremediation experiments was obtained 

from a depth of 580-590 feet.. .............................................................................. 3-5

Figure 3-4 Column experimental apparatus illustrating (A) column and effluent 

collection, (B) effluent tubing with gas purged jacketing to minimize oxygen 

invasion and effluent collection bottle, (C) influent reservoir with gas mixture 

purge and outlet lines, (D) influent lines with gas jacketing, (E) peristaltic 

pump in purged housing. F) Column L2 after 46 days of biostimulation 

showing dark banding forming in sediment (arrow) and Sterivex filter on 

outlet for microbial assay test in gas jacketed housing upstream of effluent 

collection bottle..................................................................................................... 3-13

Figure 3-5 ISR sediment uranium batch uptake experiment 1. Dissolved uranium (μM) 

versus time for different electron donor amendments. Data points are for 

each batch bottle duplicate. Control has no added electron donor or 

dissolved U(VI). The “No electron donor” bottle has 20 μM dissolved 

uranium, but no added electron donor................................................................. 3-27

Figure 3-6 ISR sediment uranium batch uptake experiment 2. (A). Dissolved uranium as 

ratio of initial total dissolved (20 μM) and (B) dissolved iron versus time for 

different headspace H2 partial pressure amendments as electron donor, and 

headspace pCO2 to control pH. Data points are for each batch bottle 

duplicate. Control has no added electron donor or dissolved U(VI). No 

electron donor bottle has 20 μM dissolved uranium added, but no hydrogen 

added to headspace. IR is gamma irradiated. ...................................................... 3-30

Figure 3-7 Columns L1 and L2 Effluent Concentrations. Part A. Column L1 and L2 

effluent dissolved uranium and iron concentrations versus days since start of 

lactate addition for the duration of suboxic elution of column L2 (top), and 

expanded time scale for biostimulated reduction (bottom)................................. 3-34
x 

Figure 3-8 Cumulative loading of uranium on column L1 and L2 sediment versus time of 

the experiment. U loading during biostimulation (Ured) is calculated by the 

difference between the influent and effluent uranium times sample volume 

and divided by total dry mass of sediment. Loading is summed up to a given 

sampling time ........................................................................................................ 3-38

Figure 3-9 Column L2 effluent dissolved uranium (red triangles) and iron (blue 

diamonds) during suboxic elution versus days since start of lactate addition. 

This is an expanded scale for Fe and time from Figure 3.7A. Effluent samples 

were collected with a fraction collection throughout the suboxic elution 

except for the period between 97 and 132 days, when samples were 

collected into serum bottles with outlet lines fully jacketed with equilibration 

gas. Oxygen (0.5%) was added to equilibration gas at day 133. Outlet line to 

fraction collector was jacketed with equilibration gas starting at day 138. 

Low Fe in effluent between days 133 and 138 was likely because of oxidation 

and precipitation in outlet line. The pump housing with equilibration gas 

flush was installed at day 153. .............................................................................. 3-41

Figure 3-10 Column L5 suboxic AGW elution of the ISR 3712 B-C 580-590 sediment. 

Effluent dissolved uranium (blue diamonds) and iron (red triangles) versus 

days. Column L1 dissolved U (green diamonds) during biostimulated 

reduction with 20 μM U(VI) influent is shown for comparison............................. 3-43

Figure 3-11 Solid phase uranium concentrations of sediments recovered after 

biostimulated reduction period (column L1) and after suboxic elution 

(column L2 and L5). (A) Uranium concentrations in columns L1, L2, and L5 by 

interval with pre-column sediment shown for comparison. (B) Mass 

distribution of uranium uptake by column interval in μmoles for columns L1 

and L2 as the increase above initial total U of pre-column sediment. (C) 

Fraction of total uranium uptake in column by interval for columns L1 and L2... 3-49

Figure 3-12 Extractable iron concentration in column sediments by interval for (A) 

column L1after biostimulated reduction and (B) column L2 after suboxic 

elution, with pre-column sediment shown for comparison. HCl represents 

ferrous iron as defined by 1 hour 0.5 N HCl extraction in absence of oxygen, 

and HH represents both ferrous and ferric as defined by 1 hour room 

temperature hydroxylamine hydrochloride extraction. ....................................... 3-51

Figure 3-13 (A). Total sulfur concentration in column sediments by interval for column L1 

after biostimulated reduction and column L2 after suboxic elution, with precolumn sediment shown for comparison. (B). S isotopic composition of total 

sulfur expressed as parts per thousand δ34S (‰).................................................. 3-54

Figure 3-14 U LII XANES spectra of column sediments after biostimulated reduction with 

lactate (L1) and suboxic elution (L2). ISR pre-column background sediment, 

and after suboxic elution of the precolumn sediment (L5) are shown for 

comparison. Numbers (e.g R1) refer to column subsection interval with lower 

numbers closer to inlet. Sample spectra are shown in solid lines and fits in 

open symbols. Models for crystalline U(IV) (uraninite, UO2) and U(VI) 

(andersonite), are shown for comparison............................................................. 3-56

Figure 3-15 (A) U LII k3

-weighted EXAFS spectra and (B) Fourier transformed radial 

distribution of EXAFS spectra (not corrected for phase shift) for ISR column 
xi

sediments after biostimulated reduction with lactate (L1) and suboxic elution 

following biostimulated reduction (L2). Pre-column experiment sediment and 

the crystalline urananite model are shown for comparison. Letters and 

numbers (e.g. R1) refer to column subsection interval with lower numbers 

closer to inlet. Uranium concentrations listed are total concentrations 

measured by hot nitric acid dissolution and include pre-column background 

U. ........................................................................................................................... 3-59

Figure 3-16 ISR pre-column sediment thin section: (A) Optical scan; (B) meso-scale XRF 

tri-color map of whole thin section panel at 30 x 30 um resolution with U 

shown in red, Fe in green, and Ca in blue. U is shown as difference between 

fluorescence above and below the U LIII edge to remove contribution from Rb 

fluorescence. (C) Area of interest for BL2-3 XRF mapping bounded by yellow. 

Scale bars are in units of microns.......................................................................... 3-61

Figure 3-17 X-ray microprobe XRF maps of region 1 and 2 for ISR pre-column sediment 

thin section shown in Figure 3.16. (A) μXRF map of region 1 at 5 x 5 μm 

resolution with iron shown in green, uranium in red and calcium in blue. 

Color brightness increases with concentration. (B) Region 2 scanned at 10 x 

10 μm resolution at a single energy with the image shown for iron in green, 

total U in red and calcium in blue. Images are mirror of Figure 3.16C and 

rotated 90° counter clockwise. Yellow circles mark grains where U LIII

μXANES spectra were attempted. The spectrum for the circled area in panel 

B is shown in Figure 3.18 (ISR Pre-column pt 1). Scale bars are in units of 

microns.................................................................................................................. 3-62

Figure 3-18 U LIII μXANES spectra of points on thin sections of column sediments after 

biostimulated reduction with lactate (L1), suboxic elution (L2) and ISR precolumn background sediment. Letter and numbers (e.g R1) refer to column, 

subsection, interval with lower numbers closer to inlet (e.g. Ox1), thin 

section panel (B =bottom; T = top), grain, and point on grain within each thin 

section. Sample spectra are shown in solid lines and fits in open symbols. 

Models for U(IV) (crystalline uraninite, UO2) and U(VI) (andersonite, uranyl 

carbonate) are shown for comparison. Sample IDs and linear combination fits 

of percentage U(IV) and U(VI) are shown in Table 3.10. Locations of point 

XANES data collection are shown in Figures 3.17, 3.20B, 3.20E, 3.23A, 3.24A, 

3.24B, 3.26, 3.28C,and 3.30................................................................................... 3-63

Figure 3-19 Column L1-R1 bottom panel of thin section: (A) Optical scan; (B) meso-scale 

XRF tri-color map of whole thin section panel with U shown in red, Fe in 

green, and Ca in blue, with areas of interest bounded by yellow. Scale bars 

are in units of microns. (C) is zoom of area 2 of meso-scale XRF map in B, 

with grains of interest denoted (e.g. 2A). (D) is a higher zoom of large iron 

grain (2A) in panel C bounded by yellow............................................................... 3-65

Figure 3-20 X-ray microprobe U oxidation state maps of grains 2A, 2B, and 2C in lactate 

column sample L1-R1 thin section shown in Figure 3.19. Panel A is a U 

oxidation state map of grain 2A where iron is shown in blue, U(VI) in green, 

and U(IV) in red. Panel (B) shows only U(IV) and U(VI). Total counts for small 

grain in upper left of 2A are 391 for U(IV) and 634 for U(VI). The total counts 

from the large grain are 5347 for U(IV) and 5962 for U(VI). Panel (C) is an iron 
xii

μXANES map illustrating distribution of iron sulfide (red), iron oxide (green) 

and S (blue). These are mirror images of Figure 3.19 and rotated 90° counter 

clockwise. Points in panel B depict locations of U μ-XANES spectra shown in 

Figure 3.18 (L1-R1 B G2 pt 1, 2, and 3), and in panel C locations of Fe μXANES spectra shown in Figure 21 (L1-R1 g2 pt 1 and 2). Panels (D) and (E) 

are U oxidation state maps of smaller grains (2B and 2C) shown in enlarged 

area of Figure 3.19C, with U(IV) in red, U(VI) in green, and Fe in blue. Total 

counts are 2151 for U(IV) and 3252 for U(VI) for the area imaged in D, and 

5850 for U(IV) and 6918 for U(VI) in E. The U μ-XANES spectrum collected at 

the center of grain 2C is shown in Figure 3.18 (L1-R1 B G2c). Scale bars are in 

units of microns..................................................................................................... 3-66

Figure 3-21 Fe XANES spectra of points on ISR column L1 sediment thin sections, and iron 

model reference mineral spectra. Points are shown in Figures 20C and 23B....... 3-67

Figure 3-22 Column L1-R2 top panel of thin section: (A) Optical scan and (B) meso-scale 

XRF tri-color map with U shown in red, Fe in green, and Ca in blue. Area of 

interest is bounded by yellow and is enlarged in C. Bottom panel of thin 

section L1-R2: (D) optical scan, and (E) meso-scale XRF tri-color map of whole 

thin section panel with U shown in red, Fe in green, and Ca in blue. Area of 

interest (grain 3) is bounded by yellow. Scale bars are in units of microns.......... 3-69

Figure 3-23 X-ray microprobe maps of thin section L1-R2 top panel, grain 1 of crescent 

shaped iron grain shown in Figure 3.22C, but rotated 90 counter clockwise, 

and mirror image. (A) is tri-color U oxidation state map U(IV) shown in red, 

U(VI) in green, and Fe in blue. Total counts in image area 12209 for U(IV) and 

16374 for U(VI). (B) is iron phase map with iron sulfide shown in red, 

ferrihydrite (iron oxide) in green, and sulfur in blue. Sulfur and iron sulfide 

map together as purple. Points in A depict locations of U μ-XANES spectra 

shown in Figure 3.18 (L1-R2 T G1 pt 1, 2, and 3), and in B locations of Fe μXANES spectra shown in Figure 3.21 (L1-R2 T g1 pt 1, 2, 3, and 4). Scale bars 

are in units of microns........................................................................................... 3-70

Figure 3-24 (A) X-ray microprobe U oxidation state map of L1-R2 bottom panel grain 3 in 

Figure 3.22E with U(IV) shown in red, U(VI) in green, and Fe in blue. Circle 

denotes the location of U μXANES spectrum L1-R2 B G3 shown in Figure 3.18. 

Total counts in imaged area are 21746 for U(IV) and 28354 for U(VI). (B) U 

oxidation state map of L1-R2 top grain 1B on Figure 3.22C with U(IV) in red, 

U(VI) in green, and Fe in blue. Circle denotes the location of the U μXANES 

spectrum L1-R2 T G1b. Total counts in imaged area are 694 for U(IV) and 

3748 for U(VI). Scale bars are in units of microns................................................. 3-71

Figure 3-25 Column L2-Ox1 bottom panel of thin section: (A) Optical scan, and (B) mesoscale XRF tri-color map of whole thin section panel with U shown in red, Fe in 

green, and Ca in blue. (C) shows area of interest for BL 2-3 microprobe 

mapping bounded by yellow in B. Scale bars are in units of microns................... 3-72

Figure 3.26 (A) X-ray microprobe XRF map of region thin section L2-Ox1 shown in yellow 

rectangle in Figure 3.25C. Fe is in green, U red and Ca blue. Color brightness 

increases with concentration. Images are mirror of Figure 3.25 and rotated 

90° counter clockwise. Grains 1 and 2 are bounded by yellow ovals. (B) 

Detailed XRF map of grain 2 at 5 x 5 um resolution. U μXANES locations are 
xiii

bounded by yellow circles. U μXANES spectra for these points are shown in 

Figure 3.18 and denoted L2-Ox1 B G1 pt 1 for grain 1 in A, and L2-Ox1 B pt 2 

and 2b in B. Scale bars are in units of microns...................................................... 3-73

Figure 3-27 Column L2-Ox3 top panel of thin section: (A) Optical scan; (B) meso-scale 

XRF tri-color map of whole thin section panel with U shown in red, Fe in 

green, and Ca in blue. (C) is an enlargement of the area of interest for X-ray 

microprobe mapping bounded by yellow in B. Scale bars are in units of 

microns.................................................................................................................. 3-74

Figure 3-28 A) X-ray microprobe XRF map at 5 x 5 um resolution map of the area within 

yellow oval in Figure 3.27C thin section L2-Ox3 top, where U is shown in red, 

Fe in green, and Ca in blue. The map is a mirror image and rotated 90°

counter clockwise of the area in Figure 3.27. (B) 2 x 2 um resolution map of 

the high U grain circled in A showing U(IV) distribution in high area 

concentration increasing with color warmness (blue to red). (C) is U oxidation 

state of the same area in B, where U(IV) is in red, U(VI) in green, and Ca in 

blue. Fe was not shown because of very low concentration relative to U(IV). 

U μ-XANES spectrum was collected at center of highest U area of the grain 

and shown in Figure 3.18 (L2-Ox3 T G1 pt1). Total counts in imaged area are 

2570 for U(IV) and 1049 for U(VI). Scale bars are in units of microns. ................. 3-75

Figure 3-29 Column L2-Ox3 bottom panel of thin section: (A) Optical scan; (B) BL10-2 

meso-scale XRF tri-color map of whole thin section panel with U shown in 

red, Fe in green, and Ca in blue. Area of interest for higher resolution 

μXANES imaging bounded by yellow is enlarged in C. Scale bars are in units of 

microns.................................................................................................................. 3-76

Figure 3-30 X-ray microprobe U oxidation state maps of L2-Ox3 bottom panel grain 1, 

the circled grain in Figure 3.29C. (A) U oxidation state at 2 x 2 um resolution 

where U(IV) in red, iron is shown in green and Ca in blue. U(VI) is not shown 

because of very low concentration relative to U(IV). (B) U(IV) distribution of 

the same grain where concentration increases with color warmness blue to 

red. U μ-XANES spectrum was collected at center of highest U area in top 

part of the grain and is shown in Figure 3.18 (L2-Ox3 B G1 pt1). Total counts 

in XANES mapped area are 50078 for U(IV) and 2339 for U(VI). .......................... 3-77

Figure 3-31 (A) Column L2 backpressure at influent end of column versus time. Pressure 

is in pounds per square inch above atmospheric pressure. (B) Bromide tracer 

versus volumes of tracer influent for column L2 at the end of biostimulated 

reduction, and for column L5 with no biostimulation to represent initial 

condition of columns, plotted as the ratio of effluent to influent bromide 

concentration. Bromide elution for column L2 is plotted as one minus the 

ratio of measured bromide to the initial influent concentration.......................... 3-89

Figure 4-1 Effluent lactate (blue symbols) and acetate (red symbols) concentrations 

from column 1 (solid symbols) and column 2 (open symbols). .............................. 4-4

Figure 4-2 Effluent Fe(II) concentrations from column 1 (blue diamonds) and column 2 

(red squares)............................................................................................................ 4-5

Figure 4-3 Effluent U(VI) concentrations from column 1(blue diamonds) and column 2 

(red squares)............................................................................................................ 4-6
xiv

Figure 4-4 Effluent sulfate concentrations from column 1(blue diamonds) and column 2 

(red squares)............................................................................................................ 4-7

Figure 4-5 Lactate concentrations in column effluent: comparison of model and 

observations from columns 1 and 2...................................................................... 4-19

Figure 4-6 Acetate concentrations in column effluent: comparison of model and 

observations from columns 1 and 2...................................................................... 4-20

Figure 4-7 Fe(II) concentrations in column effluent: comparison of model and 

observations from columns 1 and 2...................................................................... 4-21

Figure 4-8 Sulfate concentrations in column effluent: comparison of model and 

observations from columns 1 and 2...................................................................... 4-22

Figure 4-9 U(VI) concentrations in column effluent: comparison of model and 

observations from columns 1 and 2...................................................................... 4-23

Figure 4-10 FeS(s) abundance as a function of distance from column entrance at the end 

of the 83-day biostimulation. (1 dm = 10 cm)..................................................... 4-24

Figure 4-11 Sorbed U(VI) abundance as a function of distance from column entrance at 

the end of the 83-day biostimulation.................................................................... 4-25

Figure 4-12 U(IV) concentration as a function of distance from column entrance at the 

end of the 83-day biostimulation.......................................................................... 4-26

Figure 4-13 Fe(II) silicate abundance as a function of distance from column entrance at 

the end of the 83-day biostimulation.................................................................... 4-27

Figure 5-1 Model grid 1. The square domain is 196 m on a side, has the same spacing in 

the x- and y-directions, and contains 25,921 nodes. Blue dots are injection 

wells, red dot is extraction well............................................................................... 5-6

Figure 5-2 Model grid 1 center area, showing small cells used in the vicinity of the wells. .... 5-7

Figure 5-3 Model grid 2. The square domain is 396 m on a side, has the same spacing in 

the x- and y-directions, and contains 76,729 nodes. Blue dots are injection 

wells, red dot is extraction well............................................................................... 5-8

Figure 5-4 Model grid 2 center area, showing small cells used in the vicinity of the wells. ..... 5-9

Figure 5-5 Top: change in hydraulic head in meters caused by pumping wells, in 

comparison to the background flow state, grid 1. Bottom: change in 

hydraulic head caused by pumping wells, in comparison to the background 

flow state, grid 2.................................................................................................... 5-11

Figure 5-6 Top: paths of particles initially placed around injection wells, and color 

contours of head change, grid 1. Bottom: paths of particles initially placed 

around injection wells, and color contours of head change, grid 2...................... 5-12

Figure 5-7 Travel time histograms for particles released from the perimeter of the 

injection wells. The travel time intervals are in days: < 10, 10 to 30, 30 to 

100, 100 to 300, 300 to 1000, and > 1000. Results are presented for the four 

combinations of flow(east to west, and southwest to northeast) and regional 

groundwater gradient (0.0005, 0.005).................................................................. 5-14

Figure 5-8 Grid 1, Gradient 5e-04 E-W, D=0, tracer concentrations over time. Green 

dots are well locations........................................................................................... 5-15
xv 

Figure 5-9 Grid 2, Gradient 5e-04 E-W, D=0, tracer concentrations over time. ..................... 5-16

Figure 5-10 Grid 2, Gradient 5e-04 E-W, D=10, tracer concentrations over time. ................... 5-17

Figure 5-11 Grid 1, Gradient 0.005 SW-NE, D=0, tracer concentrations over time.................. 5-18

Figure 5-12 Grid 2, Gradient 0.005 SW-NE, D=0, tracer concentrations over time. ................ 5-19

Figure 5-13 Grid 2, Gradient 0.005 SW-NE, D=10, tracer concentrations over time. .............. 5-20

Figure 5-14 Tracer concentrations at 1000 days for 4 base cases (D=0) and 4 dispersion 

cases (D=10), grid 2. Black lines delineate capture zones, wherein particles 

within contour line are captured by the extraction well in advective 

transport................................................................................................................ 5-22

Figure 6-1 Lactate concentration (M) distribution after 2 days of injection from 

perimeter wells........................................................................................................ 6-4

Figure 6-2 Acetate concentration (M) distribution after 2 days of injection from 

perimeter wells........................................................................................................ 6-5

Figure 6-3 Aqueous Fe(II) concentration (M) distribution after 2 days of injection from

perimeter wells........................................................................................................ 6-6

Figure 6-4 Sulfate concentration (M) distribution after 2 days of injection from 

perimeter wells........................................................................................................ 6-7

Figure 6-5 Aqueous U(VI) concentration (M) distribution after 2 days of injection from 

perimeter wells........................................................................................................ 6-8

Figure 6-6 Adsorbed U(VI) concentration (M) distribution after 10 days of 

biostimulation from injection wells......................................................................... 6-9

Figure 6-7 U(IV) concentration (M) distribution after 10 days of biostimulation from 

injection wells........................................................................................................ 6-10

Figure 6-8 FeOOH (goethite) concentration (M) distribution after 10 days of 

biostimulation from injection wells....................................................................... 6-11

Figure 6-9 FeS concentration (M) distribution after 10 days of biostimulation from 

injection wells........................................................................................................ 6-12

Figure 6-10 Adsorbed Fe(II) concentration (M) distribution after 10 days of 

biostimulation from injection wells....................................................................... 6-13

Figure 6-11 Lactate-driven phyllosilicate Fe(III) concentration (M) distribution after 10 

days of biostimulation from injection wells. ......................................................... 6-14

Figure 6-12 Acetate-driven phyllosilicate Fe(III) concentration (M) distribution after 10 

days of biostimulation from injection wells. ......................................................... 6-15

Figure 6-13 Aqueous sulfide concentration (M) distribution after 10 days of 

biostimulation from injection wells....................................................................... 6-16

Figure 6-14 Lactate concentration (M) distribution after 30 days of injection from 

perimeter wells...................................................................................................... 6-17

Figure 6-15 Acetate concentration (M) distribution after 30 days of injection from 

perimeter wells...................................................................................................... 6-18

Figure 6-16 Aqueous Fe(II) concentration (M) distribution after 30 days of injection from 

perimeter wells...................................................................................................... 6-19
xvi

Figure 6-17 FeOOH (goethite) concentration (M) distribution after 30 days of injection 

from perimeter wells............................................................................................. 6-20

Figure 6-18 Sulfate concentration (M) distribution after 30 days of injection from 

perimeter wells...................................................................................................... 6-21

Figure 6-19 FeS concentration (M) distribution after 30 days of lactate biostimulation. ....... 6-22

Figure 6-20 Aqueous sulfide concentration (M) distribution after 30 days of lactate 

biostimulation. ...................................................................................................... 6-23

Figure 6-21 Aqueous U(VI) concentration (M) distribution after 30 days of lactate 

biostimulation. ...................................................................................................... 6-24

Figure 6-22 Adsorbed U(VI) concentration (M) distribution after 30 days of lactate 

biostimulation. ...................................................................................................... 6-25

Figure 6-23 U(IV) concentration (M) distribution after 30 days of lactate biostimulation....... 6-26

Figure 6-24 Lactate concentration distribution using continuous background 

concentrations of electron acceptors at 2, 10, 20, and 30 days. .......................... 6-28

Figure 6-25 Acetate concentration distribution using continuous background 

concentrations of electron acceptors at 2, 10, 20, and 30 days. .......................... 6-28

Figure 6-26 U(VI) concentration (M) distribution at 2, 10, 20, and 30 days using injectate 

with continuous background concentrations of electron acceptors. ................... 6-29

Figure 6-27 Adsorbed U(VI) concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-29

Figure 6-28 U(IV) mineral concentration (M) distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-30

Figure 6-29 Aqueous Fe(II) concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-30

Figure 6-30 Lactate-based phyllosilicate Fe(III) concentration distribution at 2, 10, 20, 

and 30 days using injectate with continuous background concentrations of 

electron acceptors................................................................................................. 6-31

Figure 6-31 Acetate-based phyllosilicate Fe(III) concentration distribution at 2, 10, 20, 

and 30 days using injectate with continuous background concentrations of 

electron acceptors................................................................................................. 6-31

Figure 6-32 FeOOH (goethite) concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-32

Figure 6-33 FeS mineral concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-32

Figure 6-34 Aqueous sulfate concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-33

Figure 6-35 Aqueous sulfide concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors. .... 6-33

Figure 7-1 Summary of activities leading to full-scale bioremediation deployment in the 

field. Hexagons represent characterization steps. Rectangles represent 

analysis and design steps (Long et al., 2008). ....................................................... 7-10
xvii

 Tables

Table 3-1 Kingsville Dome groundwater chemistry from wells sampled August 5-6, 

2009. Major and minor cations determined by ICP-OES of filtered (0.45 μm), 

acidified samples. Anions determined by ion chromatography of filtered 

samples. The following elements are not reported since below method 

detection limits: Co, Cu, P, F (< 2E-8, 8E-8, 1E-6 M, 5E-7M). “n/m” indicates 

not measured. ......................................................................................................... 3-8

Table 3-2 Artificial groundwater composition for column and batch experiments. 

Lactate, nutrients, trace elements and vitamins added only to AGW during 

biostimulation. ...................................................................................................... 3-10

Table 3-3 Column parameters: dimensions and weights. ..................................................... 3-15

Table 3-4 Total dissolved iron in batch experiment 1 samples filtrates (<0.2 μm) for 

duplicate bottles for each electron donor. ........................................................... 3-27

Table 3-5 Integrated loss and gain from column effluent concentrations during lactate 

biostimulation. U, sulfate, and lactate losses are the difference between 

influent and effluent concentration multiplied by sample volume summed 

over the biostimulation period. Acetate produced (net) is millimoles 

transported out of the column. Acetate consumed is the difference between 

lactate consumed and effluent acetate. Total amount of U, reduced Fe, and 

total S in sediments recovered from columns are the measured 

concentrations minus pre-column sediment concentration in each 

subsectioned intervals times mass of dry sediment recovered. L2 sediments 

were analyzed following suboxic elution. Units are all millimoles except for 

uranium, which is in micromoles........................................................................... 3-37

Table 3-6 Kingsville Dome ISR 3712 B-C 580-590 sediment grain size distribution as 

determined by Coulter grain size analyzer............................................................ 3-44

Table 3-7 Mineral weight percent of Kingsville Dome ISL sediment 3712 B-C 580-590 

(pre-column) as determined by quantitative x-ray diffraction. ............................ 3-44

Table 3-8A Column sediment solid phase uranium concentrations and whole column 

mass for column L1 after biostimulated reduction, column L2 after suboxic 

elution following reduction, and column L5 after suboxic elution. ...................... 3-46

Table 3-9 Summary of uranium oxidation state from best fits of bulk U XANES spectra 

of column sediment samples. U(IV) and U(VI) are expressed as fractions of 

total uranium from linear combination fitting. UO2 only represents fits to 

sample spectra using only crystalline uraninite model. UO2 + Andersonite 

represents fit results using crystalline uraninite as U(IV) component and 

andersonite for U(VI) component. * indicates measurement of second split 

of sample in July 2012, with first measurement in March 2010 for precolumn sediment, and December 2011 for L1-R2. All XANES were collected 

on U LII edge with the exception of the pre-column sample collected on the U 

LIII edge in March 2010. χ2 represents the sum of the square of difference 

between sample spectrum and the fit of the spectrum using the fraction of 

the components listed, and is specific to individual samples. Sample spectra 

and fits are shown in Figure 3.14. ......................................................................... 3-57
xviii

Table 3-10 Summary of uranium oxidation state from best fits of individual point 

μXANES spectra on column sediment thin sections. U(IV) and U(VI) are 

expressed as fraction of total uranium from linear combination fitting of 

crystalline uraninite (UO2) for the U(IV) component and either andersonite 

(AND) or U(VI) sorbed to ferrihydrite (FHY) for U(VI) components. U LIII edge 

μXANES were collected using a 1-μm beam size at SSRL BL 2-3. χ2 represents 

the sum of the square of difference between sample spectrum and the fit of 

the spectrum using the fraction of the components listed, and is specific to 

individual samples. Sample spectra and fits are shown in Figure 20. Sample 

IDs represent column, interval section, thin section panel, grain and point on 

grain. Locations are shown on XRF maps of thin sections in Figures 3.17, 3.20, 

3.23, 3.24, 3.26, 3.28 and 3.30. * denotes U μXANES points on large iron 

sulfide grains.......................................................................................................... 3-64

Table 3-11 Microbial assay of Kingsville Dome ISR pre-column sediment sample 3712 BC 

580-59 showing microbial phylotypes, relative abundance, and closest BLAST 

hits. ........................................................................................................................ 3-81

Table 3-12 Geobacter clones from sediments recovered at the end of the lactate 

amendment batch uranium uptake experiment using the 3712 BC580-590 

sediment................................................................................................................ 3-83

Table 3-13A Sequences obtained from effluent filter collected during later stages of Fe(III) 

reduction. .............................................................................................................. 3-83

Table 3-14 Summary of electron transfer during biostimulated reduction in column L1. 

Lactate and acetate consumed are from integrated column effluent 

concentrations (see Table 3.5). Reduction of U is from integrated increase in 

sediment recovered from column L1. Sulfate reduction is the sum integrated 

dissolved sulfide in effluent and the increase in total sulfur in column 

sediments. Iron reduction is the sum integrated dissolved iron in the effluent 

and the extracted iron from column sediments (see footnotes). Manganese 

reduction is the sum of the integrated effluent dissolved Mn and the total 

sediment Mn, assuming that it is all reduced. ...................................................... 3-87

Table 4-1 Hydrologic parameters for Column 1: Darcy flux, pore velocity, residence 

time, porosity, and dispersivity. The flow rate to the column changed at day 

14; thus, the multiple values represent those before and after that time 

point. ..................................................................................................................... 4-10

Table 5-1 Hypothetical sand-sandstone aquifer attributes..................................................... 5-1

Table 5-2 Well attributes for 5-spot pattern. .......................................................................... 5-2

Table 5-3 Model domain and sensitivity parameters.............................................................. 5-5

Table 5-4 Grid testing attributes for Grid 1 and 2. .................................................................. 5-5

Table 5-5 Particle travel time statistics.................................................................................. 5-13

Table 7-1 Prioritized Information and Monitoring Parameters for Assessment of 

Bioremediation of U(VI). Adapted from Long et al. (2008)..................................... 7-7
xix

Executive Summary

In situ recovery (ISR) of uranium, also known as solution mining or in situ leaching (ISL), 

involves the delivery of an extraction solution (typically composed of groundwater with added 

oxygen and bicarbonate) to uranium deposits resulting in the solubilization of uranium minerals. 

The uranium-bearing solution can then be pumped to the surface where uranium (and other 

minerals) can be separated, processed, and packaged. A common result of ISR is the 

persistence in groundwater of elevated concentrations of uranium and other metals, in spite of 

remediation techniques such as groundwater extraction (above ground treatment e.g., ion 

exchange, reverse osmosis) and recirculation. Complete restoration to pre-operational

groundwater values is rare and it may take several years to satisfy regulatory criteria. In 

shallow uranium-contaminated aquifers, it has been demonstrated that indigenous 

microorganisms can be stimulated to chemically reduce mobile (i.e. soluble) uranium, U(VI), to 

immobile U(IV), even at sites where abiotic reducing agents were ineffective. The principal 

conclusion from this study is that in situ bioremediation(ISB) of uranium ISR sites is a potentially 

viable alternative to standard restoration approaches.

A key attribute of uranium ISR sites is that pre-ISR baseline conditions are naturally 

reducing. Thus, the biorestoration principle for uranium ISR sites is based on a return to preISR hydrologic and geochemical conditions. The potential for engineered ISB was evaluated 

with laboratory experiments using microbiologically preserved sediments from a post-ISR site at 

the Kingsville Dome operation of Uranium Resources, Inc. in Texas. Lactate was used to

stimulate indigenous microorganisms that catalyzed the reduction of U(VI), resulting in the 

precipitation of uranium from simulated groundwater flowing through column experiments. 

Analysis of sediment from the columns after the experiments showed that all uranium was 

precipitated as U(IV). In light of the potential disruption to the native microbial community by 

ISR operations (e.g., prolonged exposure of anaerobes to oxygen), the success of this

experiment was considered to be an encouraging outcome. However, the effectiveness of 

biostimulated reduction as a viable long-term restoration strategy requires control of dissolved 

oxygen to maintain the stability of the bioreduced uranium as well as the (re)established 

anaerobic microbial community. Even the introduction of groundwater with low levels of 

dissolved oxygen to a sediment column following biostimulated reductive precipitation of 

uranium resulted in the slow reversal of the process and oxidative release of uranium. 

A reactive transport model was developed that reproduced observations from the column 

experiments. The model incorporated equilibrium and kinetic reactions including biologically 

mediated terminal electron accepting processes (TEAPs) for solid phase Fe(III), aqueous U(VI), 

and aqueous sulfate; aqueous and nonelectrostatic surface complexation for Fe(II) and U(VI); 

calcite, siderite, FeS, S secondary minerals; sulfide promoted dissolution of Fe(III) minerals, and

major ion chemistry. This enabled the model to simulate the concentrations and timing of the

monitored components in the column experiments. 

A hypothetical field-scale 5-spot ISR well pattern comprised of a central extraction well 

surrounded by 4 injection wells was simulated by a model using confined aquifer geometries 
xx 

and material properties associated with sandstone uranium ore deposits. Even with the 

simplified hydrology and assumed homogeneous distribution of properties, the operation of the 

5-spot well pattern leads to nonuniform flow fields; i.e. flow rates are location-dependent. Thus, 

some parts of the aquifer in the most direct flow paths between the injection wells and the 

production well are exposed to considerably more injectate than paths with longer travel times. 

This is important when considering how injected fluids will be distributed and therefore where, 

when and how much bioremediation will take place. Furthermore, the convention of using an 

extraction pumping rate 1% more than the total injection rates of the four perimeter wells was 

found to not necessarily be adequate to prevent excursions of injectate and/or mobilized metals. 

While most of the material injected in the perimeter wells of the five-spot pattern will be captured 

by the central pumping well, higher regional gradients increase the opportunity for excursions of 

injectate beyond the capture zone. Flow patterns will be substantially more complex for large 

well-fields in naturally heterogeneous subsurface systems. 

Biostimulation was modeled using injected groundwater amended with lactate in each of the 

four perimeter wells of the 5-spot pattern. Scenarios for ISB were investigated using the 

biogeochemical reaction network observed in the column study. The quality of injected water 

significantly affected bioremediation reaction rates. One implication is that removal of terminal 

electron acceptors, such as sulfate, from groundwater that is amended with lactate and 

reinjected into the treatment zone, should decrease the amount of lactate needed over time. 

While the introduction of less carbon might reduce the potential for biofouling (which can plug 

wells and sediment making injection of more water difficult), it will also decrease the amount of 

FeS mineral produced. These minerals help maintain the low redox potential that controls the 

quality of restored water. Some U(VI) is sorbed on the sediment but the removal of uranium

from the water column geochemically causes the sorbed uranium to enter solution making it 

available for bioreduction to very low solubility U(IV). This cycle of liberating U(VI) into solution 

followed by bioreduction progressively depletes U(VI) from the solid phase of the treatment 

zone and establishes a lower redox potential for the system, with nearly complete conversion of

uranium to the reduced low solubility form. 

The interplay between variable flow rates in the subsurface and biological reaction rates

results in spatially variable distributions of uranium and other components of interest. Under the 

modeling assumptions, bioremediated areas should become stabilized within months. While 

stabilization of a site with naturally occurring regions of lower permeability may take longer, 

bioremediation of post-ISR aquifers has the potential to more effectively restore groundwater 

quality in less time than the current state of practice. This conclusion is based on previous 

studies of ISB of uranium in shallow aquifers, laboratory studies using preserved sediments 

from an unrestored ISR aquifer, and numerical simulations of a hypothetical field 

implementation with lactate biostimulation of indigenous metal reducing bacteria. Absent from 

the weight of evidence favoring biorestoration of uranium ISR sites is a well-conceived field 

study. The unique attributes of each ISR site are likely to prevent a “one size fits all” approach 

to ISB. It is therefore important to develop a sufficiently mechanistic understanding of the sitespecific processes, properties, and conditions controlling ISB. This will help determine if

biostimulation can be successful at a given site and, if it can, what approach will be the most 

effective.
xxi

Acknowledgments

Mark Pelizza and Bobby Jemison, URI, were instrumental in providing access to the 

Kingsville Dome ISR site for sampling, drilling for ISR-mined sediments, providing water quality 

data sets used in this analysis, and helpful comments on the draft report. Bill Deutsch, PNNL 

(retired), led the drilling campaign and coordinated field activities with URI. 

We thank the following NRC staff members for their contributions to this document: Steve 

Cohen and Doug Mandeville for their guidance and useful discussions, and Elise Striz who

provided valuable insights and comments in her review of this document. We especially thank

Earl Greene (USGS) and Tom Nicholson (NRC) for their help and encouragement.

We thank the following USGS scientists for their efforts on this study: Alex Blum for XRD 

analysis, Shelley Hoeft for assistance in HPLC measurements and guidance in microbial 

protocols, Steve Silva for total and isotopic composition of sulfur in sediments, Julie Kirshtein for 

microbial assay work, and Larry Miller for many discussions on microbial and biogeochemical 

processes. Sam Webb and Ben Kocar of SSRL are acknowledged for assistance in microfocused X-ray imaging. John Bargar of SSRL provided uranium reference spectra and guidance 

in XAS data collection and interpretation. Portions of this research were conducted at the 

Stanford Synchrotron Radiation Lightsource, a national user facility operated by Stanford 

University on behalf of the US Department of Energy, Office of Basic Energy Sciences. The 

SSRL Structural Molecular Biology Program is supported by the Department of Energy, Office of 

Biological and Environmental Research, and by the National Institutes of Health, National 

Center for Research Resources, Biomedical Technology Program. 

Massively parallel processing simulations were performed on the Chinook supercomputer at 

the Environmental Molecular Sciences Laboratory (EMSL), a national scientific user facility 

sponsored by the Department of Energy's Office of Biological and Environmental Research and 

located at Pacific Northwest National Laboratory. 

xxiii

Acronyms and Abbreviations

Organizational Acronyms

BER Office of Biological and Environmental Research in DOE/SC

CESD Climate and Environmental Sciences Division of DOE/SC/BER

DOE U.S. Department of Energy

EPA U.S. Environmental Protection Agency

IFRC Integrated Field Research Challenge; SBR program

NRC U.S. Nuclear Regulatory Commission

SBR Subsurface Biogeochemical Research; CESD research program

SC (DOE) Office of Science

UMTRA (DOE) Uranium Mill Tailings Remedial Action program

Terminology

% (v/v) volume percent

% (wt/v) weight to volume percent for solid in solution

μmol/g micromoles per gram of solid

[U] dissolved uranium

μM micromoles per liter

μmol micromoles

‰ parts per thousand used to express δ34S

16S rRNA clone library technique to identify microorganism phylogeny by comparison to known 

gene sequences

AGW artificial groundwater

As arsenic

AVS acid volatile sulfide

Ca calcium

cc/min gas flow rate in cubic centimeters per minute

CO2 carbon dioxide

DGGE denaturing gradient gel electrophoresis; gene sequence comparison 

technique

DNA deoxyribonucleic acid

DO dissolved oxygen

DWS drinking water standard

Eh redox potential

Fe iron

Fe(II) iron in the +2 oxidation state, ferrous iron

Fe(III) iron in the +3 oxidation state, ferric iron

FeCO3 siderite

FeS iron sulfide mineral

FeS2 pyrite

g/L mass of solid in grams per liter of solution

GPR ground penetrating radar

GW

ISB

groundwater

in situ bioremediation

ISI in situ sediment incubator
xxiv

ISL in situ leach

ISR in situ recovery

IX ion exchange; also targeted cation removal by sorption to resin beds

Kd adsorption partition coefficient

KeV kiloelectron volts

meq/L milliequivalents per liter

Mg magnesium

mL milliliters

MLS multilevel sampler

mM millimoles per liter

mmol millimoles

mmol/g millimoles per gram of solid

Mn manganese

Mn(II) manganese in the +2 oxidation state

Mn(IV) manganese in the +4 oxidation state

Mo molybdenum

mRNA messenger ribonucleic acid

N2 nitrogen gas

N2O nitrous oxide gas

NA not applicable

O2 oxygen gas

ORP oxidation-reduction potential

Pb lead

pe negative log10 of the electron activity

pH negative log10 of the hydrogen ion activity

PLFA phospholipid fatty acid; analysis for microbial community structure 

using fatty acid biomarkers

Ra radium

rDNA ribosomal deoxyribonucleic acid

RNA ribonucleic acid

RO reverse osmosis

rRNA ribosomal ribonucleic acid

S sulfur

Se selenium

Sp. biological species

TDS total dissolved solids

TEA terminal electron acceptor

TEAP terminal electron accepting process

U uranium

U(IV) uranium in the +4 oxidation state

U(VI) uranium in the +6 oxidation state

V vanadium

XRF X-ray fluorescence

δ34S sulfur-32 isotopic composition as difference from standard reference 

material

μM

UO2

micromoles per liter

uraninite
1-1 

1 Introduction

1.1 Demand for Uranium

In 2010, nuclear energy from 104 reactors accounted for 19.2% of the electricity generated 

in the U.S. (Nuclear Energy Institute). The U.S. currently has the highest uranium requirements 

of any country, more than double that of the next country, France (World Nuclear Association). 

This is in spite of the fact that the last nuclear power plant built in the U.S. was Tennessee’s 

Watts Bar 1 in 1996. 

Uranium is acquired as uranium oxide (U3O8) concentrate, which is the end product from the 

mining industry. Over the last 5 years, the annual U.S. mine production of uranium has 

remained relatively stable at ~2000 tonnes of U3O8. This is a small fraction of the U.S. 

requirement, compelling the U.S. to import the bulk of its uranium. In 2011, U.S. civilian 

nuclear power reactors purchased ~25,000 tonnes U3O8 (equivalent) at a weighted-average 

price of $55.64 per pound, an increase over 2010 when ~21,300 tonnes U3O8 were purchased 

at a weighted average price of $49.29 (Figure 1.1). This is significant in the aftermath of the 

March 11, 2011 disaster at the Fukushima Daiichi Nuclear Power Plant when the Japanese and 

German governments announced plans to phase out nuclear power generation. To some 

degree, new and planned reactor construction in China, Russia, and India are offsetting the 

potential loss in demand. 

Figure 1.1 

Weightedaverage price of 

uranium 

purchased by 

owners and 

operators of U.S. 

civilian nuclear 

power reactors, 

1994-2011 

deliveries.
1-2 

1.2 Uranium Mining in the U.S.

Uranium is ubiquitous in the subsurface with a general crustal abundance of about 3 ppm. 

Uranium ore deposits at in situ recovery (ISR) sites typically exceed 500 ppm. The general 

requirement for uranium recovery is to find an ore body that is sufficiently concentrated and 

recoverable to economically justify exploitation. Some of the advantages of ISR over 

conventional open cut and underground mining are: (1) little surface disturbance, (2) no tailings 

or waste rock, (3) economical exploitation of lower uranium content ores, (4) minimal exposure 

of workers to dust, heavy machinery, radiation and confined space hazards, (5) no large open 

cut or underground mine to rehabilitate and (6) lower capital, infrastructure and manpower 

requirements. 

Of the 6 uranium production facilities operating in the U.S. in 2011, 5 were ISR plants (Alta 

Mesa, Crow Butte, Hobson/La Pangana, Smith Ranch-Highland, Willow Creek), while the other 

(White Mesa Mill) employed conventional milling of uranium-bearing ore. ISR plants account for 

the bulk of the U.S. uranium production. Worldwide, ISR extraction has been steadily 

increasing and in 2011 accounted for nearly half of all production, led by Kazakhstan. 

1.2.1 Uranium Ore Deposits

Types of Orebodies. Uranium deposits can be found in three principal rock types: (1) igneous 

rock of hydrothermal origin, (2) Precambrian pyrite-containing conglomerates, and (3) more 

recently formed sedimentary rock. In the latter category are sandstone formations associated 

with chemically reducing conditions. The focus of this report is on uranium roll front deposits in 

saturated sandstone formations with sufficient access and permeability for ISR to be effective. 

Uranium roll front ore deposits develop through a sequence of steps beginning with 

recharge moving downward through uranium–bearing sandstone (Figure 1.2). Dissolved 

oxygen in the water percolating through the sandstone oxidizes and mobilizes the solidassociated uranium. When the aqueous oxidized form of uranium, U(VI) (e.g., UO2

++, UO2-CO3

complexes, etc.), is transported in the groundwater to a zone of lowered redox potential (i.e., 

chemically reducing conditions), oxygen is the initial terminal electron acceptor depleted via 

reduction to water. Subsequently, mobile uranium in the +6 oxidation state is transformed 

through redox reactions to the +4 oxidation state. Under most environmental conditions, this 

results in the formation of an immobile, essentially insoluble, solid uranium mineral (e.g., 

uraninite, pitchblende, etc.). The reduction of the U(VI) is generally thought to be mediated by 

microbial processes, requiring organic matter for carbon and electron donor to enrich the 

biological respiration of terminal electron acceptors that include U(VI). 
1-3 

Figure 1.2 Uranium roll front development in confined sandstone interval showing transport of 

oxidized U and its’ precipitation in the reduced zone (NRC, 1997). 

The lenticular, crescent-shaped deposits are formed at the redox interface with reducing 

conditions downgradient and oxidizing conditions upgradient. Over time, the continuous supply 

of oxidizing uranium-bearing groundwater can deplete the reduced mineral phases (e.g., pyrite) 

and organic carbon resulting in the migration of the reduction front in the direction of 

groundwater flow. The crescent tips are often strung-out, resulting in tabular blanket 

deposits. This continual advancement results in long (~100s of meters), concentrated uranium 

deposits. Depending on the number and integrity of the low permeability confining interbeds 

(e.g., shale, mudstone), a system of tabular, sometimes interconnected roll fronts can develop. 

Individual ore bodies in the sandstone lenses vary in size, but they are commonly a few tens of 

meters wide and several meters thick. For a given roll front of interest, the continuity of the 

bottom and top confining impermeable layers is key to controlling the hydraulics required by

ISR. 

1.3 In Situ Recovery of Uranium

In situ recovery (ISR), also known as solution mining or in situ leach (ISL), involves the 

delivery of a solution to uranium deposits at depth that will result in the solubilization of solidassociated uranium. In this case, the technique targets uranium-bearing roll front deposits in 

typically confined aquifers where the uranium minerals exist primarily in the reduced oxidation 

state U(IV), maintained by chemically reducing conditions (i.e., low redox potential, or Eh). 

While in the past, and currently in some other countries, harsh chemicals were used, this is no 

longer the case in the United States. Currently, lixiviant, which is the solution pumped into the 
1-4 

ore zone to extract the uranium, consists of groundwater pumped from the ore zone and then 

returned to the subsurface after adding oxygen (either as O2 gas bubbled into the water or as 

hydrogen peroxide). The oxygen added to the water converts the low solubility U(IV) of the ore 

minerals to U(VI) which has much higher solubility. At some locations, depending on the 

mineralogy, a source of carbon dioxide is also added (either as CO2 gas or as sodium 

bicarbonate). This provides carbonate ions to complex uranium, enhancing its solubility. At 

other sites there is sufficient carbonate in the ore zone that none needs to be added. The 

pregnant solution containing the extracted uranium is then pumped to the surface where 

uranium (and other minerals) can be separated, processed, and packaged. In a typical ISR 

mine unit, hundreds of wells are drilled approximately 30-40 m apart in a “grid pattern” over ore 

bodies found in sandstone groundwater aquifers (Figure 1.3). 

 

Figure 1.3 Schematic diagram of typical 5-spot and 7-spot injection/production well patterns 

(NRC, 1997).
1-5 

1.3.1 Requirements for ISR 

Uranium ISR extraction is most feasible when the following conditions are met.

1) The formation hosting the uranium mineralization is water-saturated in an aquifer 

confined above and below by relatively impermeable strata (e.g., shale or mudstone). 

The confining layers allow more hydraulic control and containment. ISR has been 

successfully employed in partially saturated aquifers (e.g., Hobson Project, Texas) but 

this is not a common scenario for a roll front deposit because of potential issues with 

oxygenation, drawdown, and sensitive water resources. 

2) The permeability of the host sandstone unit is sufficiently permeable, > 4E-13 m2

 (0.4 

Darcy, 0.3 m/d hydraulic conductivity equivalent), to permit lixiviants to be introduced 

and circulated using a system of injection and pumping wells. This is the upper end of 

the sandstone permeability range.

3) The uranium ore is accessible to the circulated lixiviant and in a chemical form that is 

leachable by the lixiviant.

1.3.2 Lixiviants for ISR 

Uranium can be liberated into solution through dissolution of the uranium-bearing solid 

matrix using: (1) acid (e.g., sulfuric acid), (2) oxidation of the U(IV) minerals (e.g., uraninite) 

using an oxidant (e.g., oxygen, hydrogen peroxide, sodium chlorate, sodium hypochlorite, or 

potassium permanganate), and/or (3) alteration of the geochemistry that favors the formation of 

stable aqueous U(VI) complexes using salt solutions (e.g., sulfate, bicarbonate, carbonate, and 

ammonium). The ideal lixiviant is one that will oxidize the uranium in the ore and contains a 

complexing agent that will dissolve and form strong aqueous complexes that remain dissolved 

and interact little with the host rock (Davis and Curtis, 2007). 

 In the U.S., the lixiviants used for uranium ISR are typically comprised simply of an oxidant 

and alkaline complexant. Currently, the typical lixiviant is a low strength solution of bicarbonate 

and oxygen maintained at circumneutral pH. At some sites, it is only necessary to add oxygen 

to groundwater to form the lixiviant since reactions in the subsurface then generate sufficient 

carbonate to complex uranium. The oxygen is incorporated into the lixiviant by dissolving 

oxygen gas into the groundwater using above ground saturators or by sparging at depth in the 

injection well. The pH is primarily used to control calcium solubility, which affects secondary 

mineral formation. In the past, the difficulty of restoring groundwater quality to an acceptable 

condition when ammonia-based lixiviants were used led to a shift to oxygen and sodium 

bicarbonate- or carbon dioxide-based leaching chemistry by the early 1980s (Tweeton and 

Peterson, 1981). In addition to environmental issues, another concern with the use of strong 

acid or base solutions is pore clogging from secondary mineral formation (Montgomery, 1987; 

Nigor et al., 1982; Mudd, 1998).
1-6 

It should be noted that acid can be an effective, albeit non-selective, lixiviant; especially 

where carbonate minerals are in relatively low abundance. For example, ISR in Kazakhstan 

generally employs high acid concentrations without oxidants in their lixiviants. In the U.S., 

however, the use of oxygen/bicarbonate lixiviants is preferable to the regulatory agencies and 

the public, primarily for environmental considerations.

1.3.3 ISR Life Cycle

After the initiation of lixiviant introduction and the first arrival of elevated uranium 

concentrations at the production wells, it does not take very long (~days) for the concentrations 

to peak (typically 300 – 600 mg/l). This is followed by a rapid decline that slows as 

concentrations reach 25 - 70 mg/l, which is the general production range (Schmidt, 1987). The 

lower end of the range is tolerable at higher sustained pumping rates. When uranium 

concentrations decrease to 10 - 20 mg/l, the economics of an ISR operation typically dictates an 

end to the extraction phase. This may take 8 months to 3 years, with most uranium ISR 

operations ending after less than 2 years. Schmidt (1987) stated that 86% of the uranium in the 

Ruth (Wyoming) ore zone was recovered during an 11-month extraction of the subsurface with 

sodium bicarbonate solution using oxygen as the oxidant. Dissolved uranium concentrations 

peaked at 130 mg/l (as U3O8) after 3 months of leaching and steadily declined thereafter to 56.3 

mg/l after 11 months. Figure 1.4 shows a conceptual representation of the ISR process and 

associated facilities.

Figure 1.4 Conceptual representation of the ISR process, operations, and facilities (illustration 

reproduced with permission from the World Nuclear Association). 
1-7 

1.3.4 ISR Impacts and Containment Issues

ISR targets ore bodies in aquifers which lie beneath the water table. This raises some 

potential environmental issues. Leaching solution that escapes from the leach area can 

contaminate the extraction zone periphery with leached-out contaminants. The principal issues 

associated with ISR uranium extraction performance include (1) controlling the lixiviant transport 

and reactions in naturally complex physically, geochemically, and biologically heterogeneous 

subsurface materials; (2) restoring groundwater to pre-operational water quality, and (3) 

disposing of large volumes of wastewater and solutions. Post-ISR groundwater is often

characterized by significantly elevated concentrations of a number of regulated parameters. 

Redox sensitive elements are especially likely to be present at elevated concentrations; these 

include uranium, arsenic, selenium, molybdenum, and vanadium. Other parameters that have 

been observed to not return to their baseline levels include: chloride, conductivity, calcium, 

sodium, total dissolved solids, and radium-226. Thus, the primary concerns are for the 

excursion of the regulated contaminants into adjacent groundwater resources (Lands and 

Council, 1999) and the incomplete or ineffective restoration of the ISR-impacted aquifer. The 

design of an ISR operation for a targeted uranium ore body is mandated by regulation to include 

the containment of the introduced lixiviants and mobilized contaminants (e.g., Mo, As, Se, V, 

Ra-226) as well as the mobilized uranium. The objective is to prevent contamination of 

groundwater away from the ISR-impacted zone by siting and designing the operation of injection 

and pumping well networks to control solution migration and prevent unintended excursions. 

Furthermore, monitoring wells outside the treatment zone are used to detect potential 

excursions of injectate or mobilized contaminants. This includes vertical excursions into 

overlying and underlying formations that may result from casing failures and improperly sealed 

exploration bores (Marlowe, 1984; Staub, 1986). The likelihood for this occurring increases with 

the number and age of wells (Marlowe, 1984).

Water-quality effects within the well field during ISR are caused primarily by chemical 

reactions between the oxygen in the lixiviant and the geologic medium containing the uranium 

ore (Davis and Curtis, 2007). Numerous chemical interactions are possible between the lixiviant 

and the uranium ore, associated secondary minerals, and host rock formation. Common 

radioactive constituents that may be mobilized by uranium ISR activities include uranium, 

radium, radon, and their respective daughter products. Trace elements of concern with respect 

to water quality include arsenic, vanadium, zinc, selenium, molybdenum, iron, and manganese 

(Kasper et al., 1979). 

1.4 Restoration of Uranium ISR Sites

1.4.1 General Description

Once the economically extractable uranium has been removed in 1-3 years and ISR 

operations cease, the residual concentrations of uranium and other groundwater contaminants 

are known to persist at concentrations in excess of regulatory limits. The process introduces a 

considerable amount of oxidant to the area. The general issue is the recalcitrance of the 

oxidized contaminants to return to their less mobile lower oxidation states. Ideally, the 

restoration of the groundwater quality at an ISR site would simply be the reversal of the 
1-8 

oxidation and complexation processes engineered by the delivery of lixiviants to the ore body. 

However, the use of chemical reductants to lower the redox potential of the groundwater has not 

generally been effective (Schmidt, 1987). 

To achieve restoration, constituents added to the groundwater for extraction and those 

mobilized during the extraction process must be removed or rendered immobile. In some cases, 

it may also be necessary to chemically treat the geologic formation in order to reverse or inhibit 

reactions initiated during the extraction phase. The optimum restoration technique for a given 

site will largely be dictated by the inherent geologic, geochemical, and hydrologic conditions of 

that site, supported by operational observations. 

Early proponents (Buma, 1979) argued that natural geochemical processes within aquifers 

could restore ISR-contaminated groundwater. The assumption was that reductants present in 

the post-ISR sediments would lead to precipitation of reduced compounds; scavenging of 

metals by pyrite, organic matter, calcite and ferric oxyhydroxides; and adsorption by quartz, 

feldspars, and clays. While this may be a plausible scenario, it is difficult to predict the time 

required and the rates at which natural geochemical processes could attenuate the mobilized 

contaminants (Rojas, 1987), let alone answer the question whether reducing conditions would 

ever return via natural processes. 

In fact, natural attenuation, alone, has not been demonstrated to be effective in regulatory 

time frames; consequently, intervention is required to restore the groundwater quality to 

acceptable levels. During the restoration phase the concentrations of regulated elements 

mobilized during leaching must be decreased below restoration targets. The primary regulatory 

restoration goal is to return the geochemistry to (or very near) pre-operational water quality. 

Failing this, it is possible to apply for alternate concentration limits, which are above baseline 

concentrations but judged by the regulator to be protective of human health. Thus, numerical 

values of measured, pre-operational groundwater parameters (typically ~35) determine the 

primary restoration levels. 

The restoration of an ISR site typically requires long-term operation of pumps and may 

require much longer duration than the uranium extraction period. In this regard, the 

decommissioning experience at ISR uranium production facilities indicates that, in general, 

groundwater restoration represents a significant portion (approximately 40%) of the total costs 

of decommissioning (Davis and Curtis, 2007). The major cost of groundwater restoration 

activities is directly related to the volume of water pumped from or recirculated through the ore 

zone aquifer.

Recent surety bond estimates for ISR facilities indicate that the groundwater restoration 

portion of the total costs of decommissioning are higher than the approximately 40% shown in 

the 1994 data. The surety estimate for the Highland uranium project was $10.5 million out of the 

$15 million subtotal without overhead or contingency (70%) (PRI, 2006a), while the surety 

estimate for Smith Ranch was $11 million out of $14.3 million (77%) (PRI, 2006b).

Consequently, post-ISR restoration is an important part of the ISR license application 

process administered by the NRC. Specifically, the applicant must provide technical and 

financial assurances that the site can and will be restored to levels that are acceptable to the 
1-9 

regulating agencies. If the potential restoration of groundwater quality after leaching cannot be 

demonstrated, the commercial license will not be granted.

1.4.2 Restoration Processes and Phases

Restoration is based on techniques that attempt to remove residual lixiviant and lower 

aqueous concentrations of uranium and other contaminants that became elevated during ISR 

operations. Three basic methods can be applied to achieve restoration:

Groundwater Sweep. In the initial phase of restoration, contaminated groundwater is 

continuously pumped from selected wells in the ISR well field without recirculation. The intent is 

to remove mobile contaminants from areas that have been affected by the lixiviant during ISR 

and draw in uncontaminated native groundwater from outside the leach zone to replace the 

pumped groundwater. This method is called groundwater sweep. Typically, with respect to the 

contaminants associated with the ISR operations (uranium, chloride, radium, etc.), groundwater 

quality improves significantly during the groundwater sweep process (Schmidt, 1989; Rio 

Algom, 2001). One issue, however, is that large amounts of solution have to be disposed of in 

water storage facilities, evaporation ponds, sprayed on land surface or injected into deep 

disposal wells after treatment. 

Groundwater sweep alone is typically insufficient and uneconomical for complete 

groundwater restoration of commercial-scale ISR operations. Because of heterogeneities in the 

aquifers, the fresh groundwater that is brought into the ore zone does not completely displace 

the residual lixiviant (Deutsch et al., 1985). Many pore volumes1 of groundwater would need to 

be pumped in order to reach the original baseline conditions, perhaps millions of gallons for a 

10-acre leach field. Finally, as described below, groundwater sweep may cause oxic 

groundwater from upgradient of the deposit to enter into the extracted area, making it more 

difficult to reestablish chemically reducing conditions. 

Surface Treatment and Recirculation. To minimize the storage and disposal of pumped 

groundwater and more effectively restore the groundwater to regulatory standards, it is usually 

necessary to use an above-ground treatment method to remove contamination from the 

extraction zone. In this second phase of ISR site restoration, the recovered solution from the 

well field is treated to reduce contaminant levels. The treated groundwater is then recirculated 

into the contaminated aquifer zone using the same ISR pattern of coordinated injection and 

pumping wells to displace the residual contaminants and control and isolate the zone of 

groundwater restoration. In this way, contaminated groundwater in the aquifer is continuously 

diluted and displaced by less contaminated injected solutions. Ideally, the reinjected solution 

contains only concentrations at baseline values or below because reinjected contaminants will 

prolong the procedure. Contaminants in the pumped groundwater can be removed via (1) ion 

exchange resin beds that remove the metal and metalloid cations from solution via sorption, (2) 

reverse osmosis (RO) to separate the contaminants using high pressure across a semipermeable membrane, and/or less frequently (3) electrodialysis. Reverse osmosis is the most 

 

1 Pore volume, in this context, is the groundwater volume in the ore zone. It is a reference unit 

for describing the amount of liquid circulated.
1-10

common method used to treat the contaminated groundwater, but is usually employed after a 

groundwater sweep of one pore volume to remove the highest levels of contaminants from 

solution. Alternatively, the recovery stream can be run through the ion exchange beds prior to 

being fed to the reverse osmosis treatment. This is because the well field at the beginning of the

restoration is still producing considerable amounts of uranium and other constituents (e.g., large 

uranyl-bicarbonate complex molecule, calcium carbonate), which can clog the RO membranes. 

The relatively clean water stream (RO permeate) is injected back into the formation and the high 

TDS waste water stream (RO reject) is typically sent to deep disposal wells. 

The primary advantages of utilizing continuous water treatment systems in aquifer 

restoration, or a combination of groundwater sweep followed by treatment and reinjection is 1) 

reduction in the total dissolved solids in the contaminated groundwater, 2) reduction in the 

quantity of groundwater removed from the aquifer to meet restoration criteria, 3) reduction in the 

volume of wastewater requiring disposal, 4) desorption of contaminants responding to the lower 

aqueous concentrations, and 5) the means to introduce reagents to the formation to reverse or 

inhibit deleterious chemical reactions.

Reagent Amendment. A third restoration phase can be used to introduce a reagent into the 

aquifer to control the solubility of contaminants. The general intent is to return the uranium roll 

front deposit to the chemically reducing conditions that maintained the pre-operational baseline 

water quality. This is because the solubilities of many of the metal and metalloid contaminants 

of concern (e.g. uranium, selenium, molybdenum, and arsenic) were lower under pre-ISR 

mining conditions. Examples include adding reducing agents such as hydrogen sulfide gas, 

sodium hydrosulfide, or oxygen scavengers to the recirculating water to re-establish reducing 

conditions in the ore-bearing unit of the aquifer (Cameco-Resources, 2012)(Deutsch et al.,

1985; Schmidt, 1989; Rio Algom, 2001) to precipitate trace metals, including uranium. The pH 

may also be raised (e.g., using potassium carbonate) to facilitate further removal of calcium, 

metals, and radionuclides. 

Stabilization. At the end of the groundwater treatment/recirculation phase, aquifer water is 

monitored for six months to a year according to a schedule accepted by the regulatory authority 

to ensure that regulatory standards have been met, no significant impact on the water quality in 

adjacent aquifers has occurred, and there are no trends in the water quality indicative of future

deterioration. After the regulatory agencies confirm satisfactory achievement of all restoration 

parameter goals, the restoration phase is officially complete and the stabilization phase 

commences. Post-restoration stability monitoring is critical in understanding the effectiveness of 

restoration efforts and the potential for long-term water quality impacts. 

The stabilization phase was introduced by the Wyoming Department of Environmental 

Quality after experience with ISR operations had shown a pattern of increasing concentrations 

of monitored parameters that persisted after operations and restoration (Schmidt, 1987). In 

some cases, restoration had to be restarted. There are a number of possible explanations for 

the post-restoration increase in monitored parameters. Uncontaminated groundwater from the 

oxic upgradient side of the uranium roll front may be drawn into an ore zone resulting in the 

reoxidation and re-solubilization of contaminants reduced and immobilized during restoration. 

This would also make it more difficult to re-establish chemically reducing conditions. 

Contaminants adsorbed to various mineral phases may slowly be desorbing. Restoration may 
1-11

have been achieved only in the vicinity of the sampling wells; consequently, large volumes of 

contaminated groundwater may remain between the wells or in pores with limited access to the 

bulk flow. Advection would eventually transport the contaminants residing between the wells 

and/or diffusing from the less mobile pore space into the bulk flow field, to the sampling wells. 

Subsurface heterogeneities in aquifer properties would exacerbate the persistence of monitored 

contaminants as uncontaminated groundwater may be preferentially swept through the ore 

zone in a few of the most transmissive flow paths, effectively bypassing lower permeability 

zones with residual lixiviant and elevated contaminants. 

The persistence of elevated concentrations of uranium and other contaminants during the 

restoration and stabilization phases of the ISR life cycle reflects the significant disturbance 

imparted by the lixiviant. If the return to baseline geochemistry is based on the restoration of 

the pre-mining chemically reducing conditions, then there needs to be a solid-phase reservoir of 

low redox potential to sustain those conditions. One concern is that the residual post-ISR solid 

phase reservoir of Fe(II) and sulfide has been heavily depleted by the introduced oxidants to the 

degree that restoration in regulatory time frames is infeasible. 

In general, it is not easy to determine the extent the reduced minerals are oxidized in a 

typical ISR mining operation. One possible indicator is the monitoring of dissolved sulfate during 

ISR operations. Schmidt (1989) reported that sulfate concentration at the Ruth ISR pilot-scale 

test site peaked at 280 mg/liter after 2 months of leaching and declined toward the ambient 

background concentration of 100 mg/liter after 5 months of leaching. This suggests that the 

sulfide minerals in good hydrologic contact with the groundwater were significantly oxidized 

during the extraction phase of operation. The expectation is that uraninite and other reduced 

minerals present in low permeability regions (i.e., poor hydrologic contact with the bulk 

groundwater flow) would be oxidized more slowly and less completely during the extraction

phase. The open question is whether these reduced phases will have sufficient influence to 

support a return to baseline reducing conditions.

1.4.3 Regulatory Standards for Restoration

After ISR operations are completed, NRC requires the licensee to restore water quality to 

either: pre-operational (baseline) levels, drinking water standards, or alternate concentration 

limits, which are above baseline concentrations but judged by the regulator to be protective of 

human health. As of 2009, no NRC regulated ISR site (11 well fields at three sites and 34 early 

licensed R&D facilities) had been documented to have restored all groundwater constituents of 

concern within the extraction area to pre-operational background conditions (NRC, 2009b, NRC, 

2009c), which is the primary regulatory standard for restoration as defined in 10 CFR Part 40 

Appendix A Criterion 5B(5). In a survey of 77 ISR well fields in Texas, Hall (2009) concluded 

that all had received amended restoration goals for at least one regulated constituent after 

operators expended a reasonable degree of effort to restore groundwater following established 

guidelines, as determined by the Texas Commission on Environmental Quality. 

In RIS 2009-05 (NRC, 2009d), the NRC recently affirmed that all restored aquifers at NRC 

licensed ISR sites must meet the groundwater quality standards in 10 CFR Part 40 Appendix A 
1-12

Criterion 5B(5). Criterion 5B(5) states: “At the point of compliance, the concentration of a 

hazardous constituent must not exceed— 

(a) The Commission approved background concentration of that constituent in the ground 

water;

(b) The respective value given in the table in paragraph 5C if the constituent is listed in the 

table and if the background level of the constituent is below the value listed; or

(c) An alternate concentration limit established by the Commission.”

1.5 Bioremediation

In situ bioremediation (ISB) has been proposed and, in some cases, tested for the 

restoration of uranium ISR sites. ISB involves the stimulation of indigenous microorganisms 

that catalyze the destruction (of organic contaminants) or immobilization of targeted inorganic 

contaminants. Many of the post-operational contaminants found at ISR sites (e.g., U, V, Se, 

Mo) are in oxidized form as a result of the circulated lixiviants. In this case, ISB is directed at 

the conversion of mobile redox sensitive components in oxidized states to their chemically 

reduced immobile form. 

The elevated concentrations of uranium and other mobilized metals during ISR are primarily 

the result of the manipulation of the subsurface chemistry to conditions that thermodynamically 

favor their solubility. The persistence of these contaminants in spite of restoration approaches 

based on extraction (e.g., pumping) and above ground treatment (e.g., ion exchange, reverse 

osmosis) underscore the importance of (re)establishing a chemical regime similar to the pre-ISR 

baseline, which does not favor the mobilization of these metals. Unless ISR operations have 

irreversibly altered the subsurface geochemistry, the concept of returning the system to a 

previous redox state where metal concentrations are below actionable levels may be more 

tractable.

Yet, attempts to lower the redox potential using inorganic chemical reductants have not 

been observed to significantly accelerate the lowering of aqueous uranium concentrations

(Schmidt, 1987). The suggestion here is that the lowering of redox potential, while necessary to 

prevent oxidation and mobilization of sparingly soluble reduced metal phases, may not be 

sufficient to achieve the reduction of uranium and possibly other redox-sensitive metals, even 

when the reduction is thermodynamically favored. Recent field studies (Williams et al., 2011) 

have demonstrated that microbial mediation is often necessary to enzymatically catalyze 

uranium reduction in natural environmental systems even in the presence of Fe(II) and sulfide 

(i.e., abiotic uranium reduction is negligible). Also, complexed uranium in solution seems to 

require lower redox potential than the free ion in order to be reduced. From this perspective, 

bioremediation may be more efficient and effective at reducing metal contaminants, as well as 

lowering the redox potential, than chemical reductants.

ISB has been tested in shallow aquifer systems (e.g., Rifle and Oak Ridge IFRCs) with 

some success (Anderson et al., 2003; Vrionis et al., 2005; Williams et al., 2011; Wu et al., 

2006b). Insights from this body of work can be useful to uranium ISR. Conceptually, one 
1-13

potential advantage of applying ISB to ISR sites is that the goal is to return the system to a preISR redox condition characterized by very low dissolved oxygen that thermodynamically favored 

low contaminant concentrations. 

It should be recognized that biostimulation can be non-specific with respect to the 

microorganisms and terminal electron accepting processes that are catalyzed. For example, 

the iron reducing bacteria (FeRB) that are responsible for the rapid and effective reduction of 

Fe(III) and U(VI) during field biostimulation experiments at the Rifle site, are eventually 

outcompeted by sulfate reducing bacteria (SRB) for the acetate electron donor (Fang et al., 

2009; Yabusaki et al., 2007; Yabusaki et al., 2011). Impacts of the SRB activity include 1) the 

bulk of the acetate electron donor is eventually supporting SRB, not the FeRB that are 

responsible for uranium reduction, 2) the oxidation of the electron donor by SRB leads to the 

production of significant amounts of bicarbonate, which enhances U(VI) desorption, 3) 

significant amounts of biomass are produced when SRB become dominant, and 4) the bulk of 

the sulfate reduced by the SRB ends up in mineral form (e.g., iron sulfides and elemental 

sulfur). Ostensibly, the enrichment of organisms that do not reduce uranium, consume most of 

the electron donor, promote U(VI) desorption, and generate materials that may clog pores is 

disconcerting. However, pore clogging in the Rifle aquifer has not been observed to be 

significant, desorbed U(VI) is more bioavailable for biologically mediated reduction, and the 

reduced sulfide and sulfur minerals provide a buffer against oxidation. While the SRB activity in 

the shallow uranium contaminated Rifle aquifer is a complication but not necessarily an obstacle

to bioremediation effectiveness, there are no guarantees that the physical, geochemical, and 

biological conditions in the deeper sandstone ISR setting will be as conducive. Furthermore, 

the disturbance of the microbial community from prolonged exposure to ISR lixiviants and the 

effect on the response to biostimulation is not well understood. 

Another insight from the study of in situ uranium bioremediation in shallow alluvial aquifers 

is that bioreduced U(IV) in the field is more recalcitrant to oxidation than in lab experiments

(Campbell et al., 2011; Sharp et al., 2011). It has been observed that biomass on the U(IV) 

mineral, uraninite, significantly inhibits dissolution/oxidation. Impurities in the groundwater may 

be incorporated into the uraninite crystal structure resulting in lower solubility. Finally, transport 

limited by the tortuous paths within and between sediment grains can create microenvironments 

that locally maintain reducing conditions. 

The attraction to ISB is the potential for more effective restoration, shorter restoration 

periods, and lower cost. The cost driver is significant because financial surety for restoration 

costs must be provided as a prerequisite to ISR operations and maintained until restoration is 

approved. The principal drawback of using ISB to restore uranium ISR sites is that it is an 

unproven technology that relies on indigenous microorganisms to 1) catalyze the reduction of 

uranium and other ISR mobilized contaminants, and 2) permanently return the ISR-impacted 

aquifer to pre-operational conditions to stabilize the immobilized contaminants left in place. 

The dearth of well-monitored field experiments to test the approach under a realistic range of 

conditions is currently a considerable limitation. Consequently, there is a need to better 

understand the engineering of the required biogeochemical reactions and the long-term 

behavior of the restored aquifer. 
1-14

1.6 Document Description

The issue of predicting and demonstrating the long-term effectiveness of uranium 

bioremediation is common to all applications of the ISB technology. This document is the third 

in a series of NUREG/CR reports that are intended to provide background and guidance on 

uranium bioremediation technology. In these reports, the complexity of uranium 

biogeochemistry during ISB is addressed by mathematical models of coupled flow, transport, 

and biogeochemical processes that incorporate knowledge from the few intensively monitored 

field biostimulation experiments that have been performed, albeit in shallow aquifer systems. 

The objective is to develop a systematic framework to better understand the impact of sitespecific material properties and conditions on uranium bioremediation. The first report, 

“Technical Basis for Assessing Uranium Bioremediation Performance (Long et al. 2007),” 

provided an overview of uranium bioremediation, including biogeochemistry principles, design 

considerations, field performance indicators and a general approach for assessing the 

performance of uranium bioremediation in the field. The second report, “Processes, Properties, 

and Conditions Controlling In situ Bioremediation of Uranium in Shallow, Alluvial Aquifers 

(Yabusaki et al. 2010),” focused on field experiments and modeling of uranium bioremediation in 

a shallow, alluvial aquifer, as well as sensitivity analyses of aqueous uranium concentrations to 

process model parameters. The study analyzed multiple electron donors (i.e., acetate, lactate, 

ethanol); multiple electron acceptors (i.e., oxygen, nitrate, iron, sulfate) in addition to U(VI); 

density effects of groundwater amendments; and enhancement of dissolved oxygen via water 

table fluctuation and surface flooding.

1.6.1 Document Purpose, Approach, and Objectives

The purpose of this document is to provide guidance on the potential effectiveness of ISB 

for ISR uranium sites and strategies for monitoring the performance of these applications.

While a few ISR sites (e.g., Smith Ranch) have attempted ISB restoration, these cases have 

generally been without comprehensive monitoring of key biological and geochemical 

parameters (e.g., 16S ribosomal RNA, DO, ORP). One exception is the dissolved hydrogen 

biostimulation field experiment (Cabezas et al., 2011) at the Kingsville Dome site, which is 

described in Chapter 2 of this report. However, this work notwithstanding, there is currently very 

little information on ISB at actual ISR field sites. 

To address the limitations of the current knowledge base of ISR behavior and the absence 

of bioremediation field studies at post-ISR uranium sites, the interim scoping approach taken 

here is to 1) build on the available geological, hydrological, and chemical data from existing ISR 

sites, 2) factor in new knowledge and insights from bioremediation field research in shallow, 

uranium-contaminated groundwater, 3) perform uranium biogeochemical reactive transport 

laboratory studies using post-ISR sediments, 4) develop a model of biogeochemical reactions 

for the column experiments, and 5) extrapolate the model reaction network to a hypothetical 

field uranium ISR setting with a standard arrangement of pumping and injection wells. Coupled 

process modeling is used to provide the platform for integrating and upscaling the behaviors 

brought about by the biostimulation. The focus is on uranium behavior and the goal is to identify 

potential issues for ISR biorestoration that should be considered in the design of field studies. 
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Issues that will be addressed include the impact of the lixiviants on the biogeochemistry, 

biostimulation injectate composition, interplay between the nonuniform groundwater flow and 

biologically-mediated reaction rates, and the effectiveness and efficiency of uranium 

bioremediation using a standard pattern of pumping and injection wells. 

1.6.2 Document Contents

After this introductory chapter, the document continues with Chapter 2, Geochemistry and 

Restoration Issues, describing the 1) geochemistry of the uranium roll front deposits, 2) 

geochemistry of uranium ISR, 3) ISR restoration issues, 4) uranium bioremediation research, 

and 5) ISR biorestoration potential. 

Chapter 3, Laboratory Studies, describes batch and column experiments performed by the 

USGS using Kingsville Dome post-ISR sediments to investigate 1) the ability of the indigenous 

microbial population for biostimulated reduction of uranium and resulting sequestration of 

uranium from the influent artificial ground water flowed through packed columns, 2) the extent of 

release of uranium in response to suboxic conditions following electron donor amendment, 3) 

the form of the sequestered uranium, and 4) changes in sediment chemistry at the end of the 

biostimulation and re-oxidation stages of the experiments. 

Chapter 4, Simulation of Uranium Bioremediation in ISR Sediment Columns, describes the 

1) uranium bioremediation modeling studies, 2) findings from the biostimulated reduction stage 

of the column experiments using Kingsville Dome ISR sediments, 3) conceptual biorestoration 

model, 4) simulator and simulation description, 5) comparison of shallow groundwater and deep 

ISR biologically-mediated processes, 6) modeling approach, 7) simulation results, and 8) key 

findings. 

Chapter 5, Field-Scale Flow and Transport Modeling, describes the 1) hypothetical ISR 

model specifications, 2) field-scale coupled process simulator, 3) flow and transport modeling 

results, and 4) findings and implications.

Chapter 6, Field-Scale Post-ISR Biorestoration Modeling, describes the 1) biorestoration 

concepts guiding the modeling, 2) modeling approach for the hypothetical biorestoration 

simulations, 3) baseline modeling results, 4) background injectate modeling results, and 5) 

summary of findings. 

The final chapter, Chapter 7, Conclusions, Findings, and Recommendations provides 1) a

summary of the major conclusions and findings, 2) recommendations including a recommended 

path forward for ISB restoration at post-ISR sites, and 3) guidance on field characterization and 

monitoring.

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2 Geochemistry and Restoration Issues

The financial impact of the typical ISR groundwater restoration phase can be seen in the 

increased surety bond estimates for recent licensing applications. (A surety bond is required by 

the NRC to ensure that sufficient funding will be available for post-ISR decommissioning costs.) 

These costs generally reflect the long duration of active restoration and stabilization required at 

ISR uranium extraction sites, sometimes more than 10 years, which is much longer than the 1- 

to 3-year period of uranium extraction. Furthermore, no ISR site has restored all regulated 

components to pre-operational background conditions (NRC, 2009c), which is the regulatory 

standard for restoration. 

Given the cost and duration of the typical uranium ISR groundwater restoration, two 

motivating questions for this report are: 

1) Why are uranium ISR sites difficult to restore?

and

2) Can bioremediation improve the efficiency and effectiveness of restoring ISR sites to 

pre-operational conditions?

2.1 Geochemistry of Roll Front Deposits

In the oxidized sandstone and groundwater upgradient of the uranium roll front, the iron 

minerals include goethite, hematite and magnetite; whereas sulfate minerals include gypsum. 

Uranium under these oxidizing conditions occurs as U(VI), the hexavalent oxidation state, in 

dissolved, adsorbed, or mineral form. As oxic groundwater approaches the upgradient edge of 

the roll front, the remaining dissolved oxygen in the groundwater is consumed in redox reactions 

via abiotic oxidation of the reduced minerals and/or microbially-mediated oxidation of organic 

carbon. In both cases, the transfer of electrons to the oxygen terminal electron acceptor yields 

water as the product of the reduction reaction. 

Microorganisms gain energy by mediating the electron transfer process, using bioavailable 

carbon as a source of electrons and substrate for growth. The oxygen-depleted zone of the roll 

front can support anaerobic bacteria that mediate other terminal electron accepting processes 

(TEAPs) if sufficient carbon and nutrients are bioavailable. In addition to U(VI), these TEAPs 

include nitrate, manganese, Fe(III), and sulfate. Consequently, the reduced uranium minerals 

[e.g., U(IV) as pitchblende and coffinite] in the roll front are associated with the reduced forms of 

these terminal electron acceptors. Fe(II), sulfide, and elemental sulfur minerals dominate the 

solid phase products. The low redox potential supported by these reduced chemical 

components provides a buffer against oxidants entering the roll front and can be sufficiently low 

to thermodynamically favor the abiotic reduction of U(VI). 

The low redox potential baseline or pre-operational conditions at ISR sites are thus 

characterized by negligible dissolved oxygen concentrations, micromolar concentrations of 

nitrate, iron, and sub-micromolar uranium concentrations. These conditions are maintained 
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primarily by the solid phase reservoir of low redox potential and the presence of bioavailable 

organic matter that provides the substrate for microbially-mediated reduction of terminal electron 

acceptors (e.g., dissolved oxygen, nitrate, Fe(III), sulfate, U(VI)). Other redox-sensitive trace 

metals of interest are vanadium, selenium, arsenic, and molybdenum. Similar to uranium, 

vanadium and selenium under oxic conditions are mobile, and form sparingly soluble minerals 

when reduced. Conversely, arsenic and molybdenum are generally immobile under oxic 

conditions and are mobilized via bioreductive dissolution of and/or desorption from Fe(III) 

minerals. The redox potential in oxidized zones is typically about +150 mV, in the ore zone it 

varies from -150 to +100 mV and can be as low as -330 mV in more highly reduced sands.

2.2 Geochemistry of Uranium ISR 

The introduction and circulation of lixiviants during uranium ISR is designed to manipulate 1) 

the redox potential to a more oxidizing state that favors the conversion of U(IV) minerals to the 

+6 oxidation state (U(VI)), and 2) the groundwater chemistry to thermodynamically favor 

aqueous U(VI) over solid-associated forms. While ISR is targeting uranium, uranium is a trace 

component of the mineral assemblage. In this sense, the lixiviant-induced reactions are nonspecific and the bulk of the consumed lixiviant is facilitating the oxidation of the reduced iron and 

sulfur minerals, leading to the formation of Fe(III) oxides and oxyhydroxides and sulfate

(Deutsch et al., 1983). Complexants are used to increase the solubility and mobility of the U(VI). 

The introduction and circulation of lixiviants during uranium ISR significantly disrupts the 

processes responsible for creating and maintaining the lower pre-operational concentrations of 

uranium, vanadium, selenium, molybdenum, and arsenic. In particular, the oxygen commonly 

used in U.S. ISR operations to oxidize the U(IV) minerals in the ore (e.g. uraninite and coffinite), 

is consumed primarily in the oxidation of other reduced minerals that comprise the reservoir of 

low redox potential. In this case, the oxidation of Fe(II) minerals leads to secondary mineral 

formation of Fe(III) oxides and oxyhydroxides. Oxidation of elemental sulfur and sulfides results 

in elevated sulfate concentrations. The extent of the oxidation depends on the lixiviant, lixiviant 

concentration and residence time, and the abundance and availability of the reduced mineral 

phases. This “non-specific” impact of the lixiviant is also responsible for the mobilization of 

vanadium and selenium, as well as major ion complexes. Aqueous concentrations of uranium 

and the other components elevated by ISR are subject to re- and co-precipitation when the

oxidizing potential is diminished (Rojas, 1987). This may occur when the transporting solution 

comes in contact with unleached, reduced sandstone. This reversal of the ISR process does not 

naturally occur under regulatory time frames for the bulk of the leached ore zone. In fact, the 

persistence of uranium and other contaminants elevated during ISR operations, in spite of years 

of restoration effort, is a strong motivation for investigating more efficient and effective 

restoration approaches. 

2.3 ISR Restoration Issues

In Chapter 1, three basic restoration methods were described for removing residual lixiviant 

and lowering aqueous concentrations of uranium and other contaminants that became elevated 

during ISR operations: groundwater sweep, above ground treatment, and reagent injection. 
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Here we describe scenarios that might explain the persistence of contaminant concentrations in 

spite of these restoration methods.

Post-ISR groundwater sweep involves pumping the ISR-contaminated groundwater out of 

the aquifer and allowing uncontaminated groundwater to be drawn in to replace it. If there are 

no other sources of contaminated groundwater, this approach alone would restore the aquifer. 

The general ineffectiveness of this approach to restore an ISR-impacted aquifer to baseline 

conditions implies that the lixiviants and/or contaminated groundwater are not completely 

removed. Residual lixiviant can react with the solid phases, continuing uranium leaching and 

contaminant production. Residual contamination from ISR extraction could be maintained in 

low permeability zones that are not accessed by the bulk flow paths. Analogously at the pore 

scale, intergranular and intragranular pore spaces with limited transport access may also 

sequester residual contaminants. These residual contaminants can become long-term 

diffusion-limited sources. 

If the residual lixivant and aqueous contaminants from ISR processes are removed by the 

groundwater sweep, then persistent contaminant concentrations are due to sources in the 

leached aquifer. Any uranium in solution can be considered to be U(VI) because of the 

extremely low solubility of U(IV). Thus, the persistence of U(VI) and oxidized forms of other 

redox sensitive metals (e.g., vanadate, selenate) implies continuing oxidation of key 

components of the reduced mineral assemblage or continuing release of oxidized forms of the 

contaminants. Continuing oxidation in spite of the groundwater flush can occur if groundwater is 

becoming oxidized through solid phase reactions or the groundwater entering the ISR-impacted

aquifer contains oxidants. 

Chemically reducing groundwater drawn in during groundwater sweep can be oxidized 

through reactions with minerals that act as oxidants. Fe(III) minerals, especially oxides and 

oxyhydroxides formed during ISR mining, can potentially provide a reservoir of oxidizing 

potential. The capacitance of this reservoir will depend on the type and concentration of oxidant 

used in the lixiviant, and the extent and duration of contact with the solid phases. 

Oxidizing groundwater can potentially enter the ISR zone of the aquifer. One scenario for 

this to occur is when groundwater is being drawn from the upgradient side of the redox interface 

of the roll front, where oxidizing conditions exist. Another scenario is oxidation of produced 

groundwater prior to reinjection. This could occur during handling such as surface treatment or 

reagent addition. In either case, the oxidizing groundwater could potentially continue the 

oxidation and release of uranium and other redox sensitive contaminants. 

Continuing release of oxidized forms of uranium and other contaminants is possible in spite 

of lowered redox potential groundwater that thermodynamically favors reduction. In this case, 

the reaction kinetics significantly limit the rate of reduction. Nonequilibrated redox couples are 

not indicative of the system redox potential (e.g., Eh, ORP). For example, U(VI) may be 

present and possibly increasing in concentration under sulfate reducing conditions even though 

the U(VI)/U(IV) redox couple is higher on the redox ladder and energetically more favorable. 

Reinjected groundwater is sometimes augmented with chemical reducing agents (e.g., 

hydrogen sulfide) to create conditions designed to remove the redox-sensitive contaminants 

(e.g., U, Se, V) from solution. This restoration technique has been used with varying degrees of 
2-4 

success (Cameco-Resources, 2009; Crow Butte Resources, 2000). At the Ruth pilot scale test 

site in Johnson County, Wyoming, six weeks of hydrogen sulfide injection lowered 

concentrations of dissolved uranium, selenium, arsenic, and vanadium by at least one order of 

magnitude (Schmidt, 1987). However, the reducing conditions were temporary as uranium, 

arsenic, and radium concentrations began to increase 1 year later. More recent studies have 

determined that natural mineral assemblages can be much more resistant to abiotic Fe(II)-

driven uranium reduction than model minerals (Jeon et al., 2005) and that abiotic sulfide 

promoted reduction of uranium can be strongly inhibited by bicarbonate (Sani et al., 2005). This 

could explain the cases where sulfide-promoted reduction of U(VI) is ineffective (Yabusaki et al., 

2011). 

2.4 Uranium Bioremediation

Biorestoration of uranium ISR sites is not a novel concept. Results from the few sites (e.g.,

Smith Ranch) that have attempted to use a bioremediation approach to restore uranium ISR 

sites have not been sufficiently successful to warrant widespread adoption. Most of these 

efforts have been relatively ad hoc approaches where the limited descriptions that are available 

do not provide sufficient detail to ascertain the biorestoration design or monitoring results. One 

notable exception that has been presented (Cabezas et al., 2011) but not yet published is a field 

biostimulation experiment at a post-ISR site in Kingsville Dome, Texas, led by Dr. Lee Clapp at 

the Texas A&M University-Kingsville. In this bioremediation design, dissolved hydrogen serves 

as the electron donor and reductant, and bicarbonate is the carbon source. Biostimulation of 

the indigenous microbial community with a 2-month pulse of dissolved hydrogen lowered U(VI) 

concentrations to the pre-operational baseline condition and have maintained those conditions 

for 2 years. Monitoring included data on 15 water quality measures. 

The U.S. DOE has sponsored field research studies to investigate in situ bioremediation of 

shallow uranium-contaminated groundwater plumes (Anderson et al., 2003; Vrionis et al., 2005; 

Williams et al., 2011; Wu et al., 2006a; Wu et al., 2006b) in much smaller field systems, on the 

order of 10 m. Intensive characterization and monitoring have provided comprehensive data 

sets and new knowledge on the biogeochemistry of uranium under engineered biostimulation of 

indigenous microoorganisms. Drilling techniques that minimize sediment sample disturbance 

without introducing additional fluids are used to recover sediments suitable for physical, 

geochemical and biological characterization and experimentation. Coupled process modeling is 

an important component of these research studies with the goal of achieving a quantitative and 

predictive understanding of the subsurface processes controlling uranium bioremediation. 

Spatially variable hydraulic conductivity, porosity, and dispersivity needed for the modeling are 

characterized using particle size distribution, density, surface area, and porosity in conjunction 

with field geophysics (e.g., neutron, gamma, resistivity logging) and aquifer flow and transport 

studies.

In these research studies, minerals and adsorbed components, especially those involved in 

active redox couples (Fe, Mn, S, U, Se, V, As), are characterized through a variety of 

techniques including point counts, extraction, x-ray diffraction, and x-ray fluorescence (XRF), as 

well as spectroscopic techniques that provide elemental associations and oxidation states. A 
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key concept is the labile fraction of the solid phase components, which addresses availability for 

chemical and biologically-mediated reactions. For example, total uranium (e.g., measured by 

XRF) in sediments at the crustal abundance level (~3 mg/L), is generally much higher than the 

carbonate extractable uranium associated with surface complexes. Similarly, the bioavailability 

of phyllosilicate Fe(III) terminal electron acceptors is greater than crystalline forms. These 

determinations required more refined analytical characterizations than bulk approaches alone. 

Groundwater from the monitoring wells is analyzed for water quality (e.g., pH, temperature, 

specific conductivity), major ions, and metals similar to the typical ISR operations. However, 

unlike most ISR operations, there is more attention to the oxidation states and speciation of the 

aqueous components. In particular, dissolved oxygen and oxidation-reduction potential (ORP) 

or Eh are monitored as well as sulfide and methane. 

A key modeling need is to be able to predict the evolving concentrations of the oxidation 

states for key redox couples under the interplay of flow, transport, and biogeochemical 

processes. To this end, the shallow uranium bioremediation field research studies employ 

laboratory-scale experiments using native sediment, groundwater, and microorganisms. These 

experiments allow detailed observation of the uranium biogeochemical dynamics under 

controlled conditions using microcosm, chemostat, batch, and column configurations. Most 

importantly, the experiments provide important insights on the dominant reactions and 

associated kinetics for various electron donor-microorganism combinations using native 

populations and aquifer materials. 

Finally, the largest departure from the typical ISR operation is the focus on the microbial 

ecology before, during, and after biostimulation. In the shallow uranium bioremediation field 

studies, molecular biological techniques are used to identify 1) microorganisms that comprise 

the native microbial community (e.g., Campbell et al., 2012), 2) microorganisms that become 

active in response to particular electron donors (e.g., Wrighton et al., 2012), and 3) specific 

metabolic responses by the active microorganisms to electron donor amendments, geochemical 

conditions, and interactions with the microbial community (e.g., Fang et al., 2012). The 

principal interest is in microorganisms that can be stimulated to catalyze useful reactions such 

as the conversion of aqueous contaminants to immobile, solid-associated form. There are other 

behaviors of interest that will also control the efficiency and effectiveness of the bioremediation 

such as microorganisms that 1) consume significant amounts of the provided electron donor, 2) 

inhibit or reverse the desired reactions (e.g., oxidizers involved in redox cycling), 3) maintain 

desirable redox conditions, 4) produce significant biomass, and 5) facilitate the continuing 

effectiveness of bioremediation. 

The efficiency and effectiveness of the biologically mediated reactions can be affected by 

ISR operations and can, in turn, affect ISR operations. Of most concern is the impact of the 

injected lixiviant on key members of the microbial community. Oxygen and other oxidants used 

to enhance leaching inhibit the growth of anaerobic microorganisms and can have toxic effects 

on strict anaerobes. The degree of disruption to the microbial community, especially to 

microorganisms that catalyze the reduction of uranium and other metal contaminants, is an 

important and open issue. Conversely, amending groundwater with carbon electron donors 
2-6 

can stimulate growth in microorganisms that leads to the accumulation of biomass in the pore 

spaces. If the biomass accumulations are sufficiently large, flow rates and flow paths can be 

altered. 

2.5 Findings from Uranium Bioremediation Field Research in 

Shallow Aquifers

The U.S. Department of Energy (DOE) has been sponsoring Integrated Field Research 

Challenge (IFRC) projects at a former Uranium Mill Tailings Remedial Action (UMTRA) site in 

Rifle, Colorado; Area 3 at the Y-12 facility in Oak Ridge, Tennessee; and the 300 Area of the 

Hanford Site in Washington. These comprehensive interdisciplinary studies are investigating 

the behavior of uranium in the subsurface environment, including the potential for field-scale

bioremediation (Anderson et al., 2003; Gu et al., 2005; Yabusaki et al., 2008). While these were 

shallow sites addressing uranium groundwater plumes, there are several findings from these 

studies that could be useful to the investigation of biorestoration of ISR sites. 

Ineffectiveness of abiotic uranium reduction under experimental time scales. 

Field experiments in the shallow alluvial aquifer at the Old Rifle former uranium mill and 

tailings site in Colorado have found no evidence of significant abiotic uranium reduction in the 

presence of elevated Fe(II) or sulfide. In these experiments, U(VI) was thermodynamically 

favored to be reduced to U(IV). The literature is somewhat equivocal, for example, uranium 

reduction has been shown by abiotic reaction on the surfaces of solid phases that form during 

biostimulation, such as iron sulfides (Hyun et al, 2012; Hua, 2008). In contrast, there is a body 

of research in shallow uranium-contaminated groundwater plumes that has observed uranium to 

be largely recalcitrant to abiotic reduction via pre-existing reduced natural mineral assemblages

in aquifer sediments. The Fe(II) behavior at the Rifle site is consistent with the Jeon et al. 

(2005) finding that abiotic U(VI) reduction on natural mineral assemblages with Fe(II) was 

negligible with the exception of Fe contents above 18%. Similarly, thermodynamically favored 

sulfide promoted U(VI) reduction has not been observed at the Rifle Site even when ~10 mM 

sulfate has largely been converted to sulfide by sulfate reducing bacteria (Yabusaki et al., 

2011). The literature is variable but the Rifle field experiments are consistent with the Sani et 

al. (2005) observation that elevated bicarbonate concentrations, not uncommon to many 

environmental situations, can limit the rate of abiotic U(VI) reduction by sulfide. This may 

explain why the restoration of some ISR sites, such as the Ruth pilot test site in Johnson, 

Wyoming benefited, at least temporarily, from injected sulfide (Schmidt, 1987) and others, such 

as the Smith Ranch-Highland site in Converse County, Wyoming (Borch et al., 2012) did not. 

Metal reducing bacteria can be stimulated to catalyze the reduction of uranium and other 

metal contaminants under experimental time scales. 

In light of the general difficulty of abiotically reducing U(VI) in shallow groundwater systems, 

it is notable that there has been considerable success using engineered biostimulation of 

indigenous microorganisms to catalyze the reduction of mobile U(VI) to immobile U(IV) as 

sparingly soluble minerals or solid associated forms. In saturated sediments, this ability has 

been demonstrated with a variety of electron donors (Barlett et al., 2012) and microorganisms 
including ethanol-oxidizing sulfate reducing bacteria at the Oak Ridge Site (Cardenas et al., 

2010) and acetate-oxidizing iron reducing bacteria at the Rifle Site (Anderson et al., 2003). This 

removal of uranium from solution can lower concentrations below actionable levels provided 

there is sufficient transport, active U(VI)-reducing microorganisms, and bioavailable electron 

donor to treat the zone of interest. 

Although U(VI) was the target of the 

bioremediation studies at the Rifle Site, 

other metals and metalloids including V 

(vanadium), Se (selenium), Mo 

(molybdenum), and As (arsenic) were 

affected by the biostimulation events 

(Figure 2.1). Vanadium was the original 

ore of interest when the Old Rifle Mill 

opened in 1924. In a recent field 

experiment in 2010, ambient V 

concentrations were reduced from 2300 

ppb to less than 20 ppb after 23 days of 

acetate amendment (Yelton et al., 2012). 

Se concentrations were reduced from 120 

ppb to less than 5 ppb after 7 days 

(Williams et al., 2012). ~15 ppb Mo in the 

form of molybdate was removed from 

groundwater down to 1 ppb at the same 

time its structural analogue, sulfate, was 

depleted via respiration by sulfatereducing bacteria. Dissolved arsenic 

concentrations increased from ~90 to ~150 

ppb during iron reduction, followed by 

larger increases up to 600 ppb once 

sulfate reduction became the dominant 

metabolic pathway and aqueous sulfide 

concentrations increased (Stucker et al., 

2012). Once acetate amendment ceased, 

the arsenic concentrations dropped below 

influent concentrations. 

Biogenic U(IV) can be in non-mineral form.

Recent laboratory studies (Bernier-Latmani et al., 2010) have shown that, depending on the

geochemical conditions, the biogenic U(IV) product may either be crystalline nanoparticulate

Figure 2.1 Downgradient metal concentrations responding 

to acetate injection at the Rifle IFRC site. Concentrations 

were measured in well CD-01 from June 2010 to May 2012. 

(Courtesy of Ken Williams, LBNL)

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2-8 

uraninite or monomeric U(IV) complexes that are associated with biomass and mineral surfaces. 

Monomeric U(IV) has been shown to be preferentially produced as a result of the interaction of 

U(VI) with phosphate (Boyanov et al., 2011; Veeramani et al., 2011). The emerging conceptual 

model for U(IV) associated with biomass appears to involve the binding of U(IV) to phosphato 

moeities on the cell biomass. Conversely, in the absence of phosphate, magnetite was shown to 

reduce U(VI) to uraninite. Thus, phosphate appears to preclude the precipitation of uraninite 

and to play a critical role in the formation of monomeric U(IV) (Bargar et al., 2013). 

Monomeric U(IV) is more labile than uraninite and this property has been used to develop

extraction techniques that can distinguish between uraninite and monomeric U(IV) (Alessi et al., 

2012). The higher susceptibility of monomeric U(IV) to oxidation is a concern for uranium in situ

bioremediation. The characterization of the stability of monomeric U(IV) complexes in aquifers

will be the subject of future applications of the recently developed extraction technique. Of 

particular interest is the sorption affinity and morphology of molecular-scale U(IV) structures 

adsorbed to various minerals of interest (e.g., alumina, silica, and montmorillonite). 

Electron donor is consumed primarily in sulfate, Fe(III), and C(IV) TEAP reactions. 

Only a small fraction of the electron donor is supporting the U(VI) and other desirable trace 

metal TEAP reactions (e.g., Se, V). Thus, the bulk of electron donor is consumed in 

biologically-mediated TEAP reactions involving the reduction of Fe(III), sulfate, and C(IV) 

terminal electron acceptors. At the Rifle Site, the metal-reducing bacteria can be outcompeted 

by sulfate-reducing bacteria for the available acetate. In this case, acetate must generally be 

provided in excess of the SRB demand to maintain U(VI) bioreduction (Williams et al., 2011). 

Products of the biostimulation (e.g., bicarbonate, biomass, Fe(II), sulfide, elemental 

sulfur, methane) can alter the hydrologic and geochemical conditions controlling 

uranium mobility. 

The products of the biologically-mediated oxidation of electron donor impart several 

changes to the hydrologic and geochemical conditions in the aquifer. Elevated bicarbonate 

resulting from the oxidation of acetate, lactate, or ethanol is thermodynamically favored to 

complex with U(VI) and other metals leading to the desorption of adsorbed forms of these 

metals. When biostimulation ceases, the vacated surface complexation sites can be 

repopulated by the adsorption of aqueous metal species. 

The reduction products (e.g., Fe(II), sulfide, elemental sulfur, methane) from the biologicallymediated TEAP reactions can participate in subsidiary reactions that lead to the formation of 

secondary minerals (e.g., FeS, calcite). While the precipitation of these minerals in situ can 

potentially incorporate uranium, the consistent observation has been for these reduced 

secondary minerals to provide a reservoir of low redox potential that can protect the aquifer 

against rapid oxidation (Abdelouas et al., 2000). Subsequent oxidation of solid-associated and 

mineral forms of Fe(II) can lead to the secondary formation of “fresh” Fe(III) mineral surfaces 

that are more reactive. The formation of secondary minerals as well as the production of 

biomass can lead to pore-clogging, which can limit the distribution of electron donor and the 

effectiveness of the biorestoration (Li et al., 2009). 
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Microorganisms controlling important TEAP reactions are subject to evolving 

geochemistry and community interactions. 

Engineered biostimulation will be a significant perturbation to the native microbial 

community. Consequently, the metabolic status of key microorganisms will evolve with 

concentration changes in electron donors and acceptors, nutrients, biomass, geochemistry, and 

interactions within the microbial community (e.g., competition, shared metabolites, population 

stress). At the Rifle Site, observed changes include community composition (e.g., Anderson et 

al., 2003), cell numbers (e.g., Chandler et al., 2010), activity (e.g., Wilkins et al., 2009), reaction 

rates (e.g., Williams et al., 2011) and attachment/detachment (e.g., Kerkhof et al., 2011). 

2.6 Insights from Field Studies

There were also some important insights from the shallow aquifer uranium bioremediation 

research studies, which can potentially be generalized to the biorestoration of uranium ISR 

sites. 

One size does not fit all. The shallow uranium bioremediation research sites have different 

site attributes that cannot be effectively addressed with the same deployment strategies. For 

example, ethanol and glucose were found to be effective electron donors at the Oak Ridge Y-12 

site, which is characterized by low pH and high nitrate, aluminum, and calcium. Under the suboxic, high alkalinity, circumneutral pH conditions at the Rifle Site, acetate was the selected 

electron donor. There were also differences in the hydrogeology that affected the electron 

donor delivery strategy. 

Natural uranium deposits are not immune to hydrologic, geochemical, and microbiological 

variability. The expectation is that effective and efficient restoration of ISR sites will be founded 

on an understanding of site-specific processes, properties and conditions. Key attributes 

include the hydrogeologic setting, dissolved oxygen, bicarbonate, major ion chemistry, terminal 

electron acceptor concentrations, and reactive mineral distribution. An important complication is 

the extent to which the lixiviant and ISR extraction has altered the hydrology, geochemistry, and 

status of the microbial community. 

Laboratory ≠ Field. In many cases, the use of model minerals, limited chemical components, 

artificial groundwater, pure cultures, and/or batch/microcosm/chemostat experiments results in 

behaviors that are not consistent with the field observations. Examples of these laboratoryscale effects include abiotic uranium reduction, uranium re-oxidation and remobilization, and 

generally higher reaction rates versus the field. Some of these behaviors observed in batch and 

column experiments can be attributed to the use of sediments from a smaller size fraction (e.g., 

< 2 mm) than the complete in situ particle size distribution. Consequently, the proportion of the 

more reactive fines is over-weighted.

Laboratory-scale studies are still an indispensable characterization activity that provides a 

framework for understanding field-scale behaviors; however, care needs to be practiced when 

extrapolating lab-scale behavior to the field. Laboratory studies should therefore include 

experiments using sediments, groundwater, and the microbial community from the targeted 
2-10

aquifer to better characterize and understand the site-specific system response to electron 

donors. 

Work within the natural biogeochemical tendencies of the system. Approaches that work 

within the processes, properties, and conditions that a given aquifer will naturally support, are 

more likely to succeed than approaches that try to impose artificial conditions. The farther the 

targeted conditions are away from the natural unmitigated state of the subsurface system, the 

more difficult it will be to achieve long-term success. For example, a uranium bioremediation 

design that relied on maintaining locally reducing conditions in an otherwise oxic aquifer would 

require considerably more effort than at the Rifle site, which is naturally suboxic. In this respect, 

returning an ISR-impacted aquifer to a state similar to the naturally reducing conditions of the

pristine uranium roll front deposit should increase the prospects for successfully engineering

biorestoration. 

Recalcitrance of bioreduced U(IV) to re-oxidation. In spite of laboratory studies that have 

identified reoxidation and remobilization of bioreduced uranium (Komlos et al., 2008a; Komlos et 

al., 2008b; Moon et al., 2007; Moon et al., 2009), recent studies (Campbell et al., 2011; Sharp et 

al., 2011) have identified recalcitrance of uraninite to reoxidation, especially under in situ

conditions that include the presence of biofilms and the full complement of geochemical 

components. One implication is that uranium reduced to U(IV) during ISR biorestoration may 

potentially withstand periods of residual oxygen exposure as the redox potential is lowered to 

near-native conditions. 

2.7 ISR Biorestoration Potential

There have been relatively few field-based studies that have been published on the cause(s) 

of the recalcitrance of uranium ISR sites to restoration. While in situ bioremediation has been 

attempted at uranium ISR sites, there are no publications of biorestoration field studies that 

have been designed and monitored to better understand the site processes, properties, and 

conditions that control its effectiveness. The need for such studies is tempered by the cost and 

complexity of comprehensively interrogating a deep aquifer with a large range of 

biogeochemical states. However, there is no substitute for well-conceived field studies of the 

uranium ore sites, alteration of those sites by ISR operations, and the engineering of biologically 

mediated reactions intended to more effectively and efficiently restore pre-operational

conditions.
3-1 

3 Laboratory Studies of Biostimulated Uranium Reduction in 

ISR Aquifer Sediments and Suboxic Remobilization of 

Sequestered Uranium (C.C. Fuller and K. Akstin)

3.1 Introduction

3.1.1 Objectives

The laboratory experimental studies with aquifer sediment recovered from an 

unconsolidated sand U-ore body previously extracted by in situ recovery (ISR) are described in 

this chapter. Flow-through column and batch experiments were designed to investigate the 

ability of the indigenous microbial population for biostimulated reduction of uranium and 

resulting sequestration of uranium from the influent artificial ground water flowed through 

packed columns. The extent of release of uranium in response to return to suboxic conditions 

also was tested on a column following uptake of uranium during biostimulated reduction. 

Analyses of sediments recovered from columns at the end of the biostimulation and re-oxidation 

stages were conducted to characterize the form of the sequestered uranium and to quantify 

changes in sediment chemistry during the course of the experiments. Biostimulation column 

experiments were conducted using an artificial ground water to represent the major ion 

chemistry of the aquifer and were continued well into sulfate reduction to simulate long-term in 

situ bioremediation effort. 

Bioremediation strategies for uranium contaminated aquifers are based on the ability of the 

microbial community to reduce the highly soluble hexavalent uranium (U(VI)) to sparingly 

soluble +4 oxidation state (U(IV) which then precipitates to form phases such as uraninite 

thereby immobilizing uranium and lowering dissolved concentrations (Lovley et al, 1991). 

Bioremediation is achieved through stimulation of the ambient microbial population through 

amendment of the aquifer with suitable electron donor to consume oxygen and promote 

anaerobic conditions. Dissimilatory metal reducing bacteria have been shown to concomitantly 

reduce U(VI) with the bioreduction of Fe(III) phases, both in the laboratory (Lovley et al, 1992), 

and field tests in shallow aquifers (Anderson et al, 2003). Sulfate reducing bacteria have also 

been shown to reduce uranium directly through enzymatic mechanisms (Lovley et al, 1993). In 

addition to direct enzymatic reduction by microbes, U(VI) reduction may also occur by abiotic 

reaction with products of biogeochemical reduction processes, such as iron sulfides (Veeramani 

et al, 2013; Hyun et al, 2012; Hua, 2008). However, the importance of abiotic, non-enzymatic 

uranium reduction has not been demonstrated at the field scale. 

3.1.2 Kingsville ISR Site Description

Aquifer sediments were recovered from the Kingsville Dome ISR facility in a roll front deposit 

previously extracted using ISR. The Kingsville ISR facility is located about 40 miles southwest of 

Corpus Christi, Texas, and about 8 miles southeast of Kingsville, in Kleberg County (Figure 

3.1). The uranium ore bodies of the Kingsville Dome area are found in the Goliad Formation, 

Miocene to Pliocene deposits of unconsolidated sand and sandstone. The Goliad is mostly 
3-2 

fluvial deposits and is part of the South Texas Gulf Coastal Plain Physiographic Province. The 

Evangeline aquifer is associated with the Goliad formation.

Uranium in the South Texas uranium belt is thought to have leached from volcanic ash 

generated by volcanic activity further west which was deposited as portions of the Catahoula 

formation (Nicot et al. 2010, Galloway, 1977). Uranium ore at Kingsville Dome is in roll front 

deposits, formed as oxidized groundwater carrying dissolved U(VI) came into contact with 

reducing conditions which are thought to have been generated by H2S bearing water migrating 

upward, around the Kingsville salt dome, from sour hydrocarbon deposits. The ore deposits are 

30 to 70 feet wide and are found at depths of 500 to 750 feet. According to the 2011 URI 

Annual report, the Kingsville dome ISR area contains proven reserves of 0.050 million tons of 

U3O8 at percent grade of 0.088%. Since July, 2009 no extraction activities have taken place on 

the 2135 acre site.

At Kingsville Dome, as described by Arrendondo (1991) from thin sections, the Goliad 

material is “fine to medium grained, calcareous quartz sand and silt with minor amounts of chert, 

K-feldspar, volcanic rock fragments, pyrite, marcasite, leucoxene, pyrite and marcasite 

intergrowths, Ca-montmorillonite, gypsum, and traces of plagioclase, kaolinite, metamorphic 

rock fragments, mica, chlorite, and barite.” While quartz comprises up to 30% of the framework 

grains, coatings of clay (primarily Ca-montmorillionite) and calcite are common.

XRD of Kingsville uranium bearing material, showed only quartz, calcite, kaolinite, and Camontmorillionite. One of the ten XRD samples showed uraninite, otherwise no U minerals were 

observed. Uranium appears to be in several forms, besides uraninite. SEM/EDAX showed U in 

the clay fraction with some being non-crystalline and adsorbed on the surfaces of thin 

authigenic Ca-montmorillonite clay coatings in quartz sand grains. It is also associated with 

detrital iron-titanium oxide minerals that have been replaced by iron disulfide minerals, pyrite 

and marcasite. Electron density microphotos of heavy minerals show Fe, Ti, U, S, and Ca with 

U commonly occurring with Ti (Arrendondo, 1991).

At the ISR site, wellfield 13, there are three stacked roll-front deposits, each about 10 feet 

thick with several feet between them. Coring was done at the north end of the site (Figure 3.2). 

Samples were taken close to borehole AA 24 in the A & B sands. Figure 3.3 shows the 

material taken from a depth of 580-590 feet that was used in the bioremediation experiments.
3-3 

Location of Kingsville Dome ISR Site

North

10 miles

After TBRC

Figure 3.1 Location of Kingsville Dome ISR site in south Texas.
3-4 

Figure 3.2 Samples were obtained by drilling into an ore zone after uranium recovery operations 

were finished.
3-5 

3.3 The sediment sample used for the bioremediation experiments was obtained from a depth of 

580-590 feet.
3-6 

3.2 Methods

3.2.1 Sediment and Groundwater Collection and Analyses

3.2.1.1 Aquifer sediment sampling

Aquifer sediments were collected August 5 to 7, 2009, from two drill holes across three 

vertically stacked roll front deposits, which all have had U extracted by in situ leaching. The 

upper and lower ore deposits (A, C) are thought to be from the reduced portions of the roll fronts 

while the middle zone (B) is from the oxidized region of the roll front. Because of near complete 

loss of sediments from core barrel during recovery, only one sediment sample, 3712 BC 580-

590, was recovered from the corer. The sediment sample was transferred directly into a Mylar 

bag with minimal handling to avoid contamination, and immediately sealed with scrubbers to 

remove oxygen. The remaining samples were recovered by washing sediment out of the bore 

hole with a high-pressure water stream. Samples were immediately sealed in Mylar/foil 

pouches containing scrubbers to remove oxygen.

Upon return to Menlo Park, the sediment samples were transferred from the Mylar 

pouches to 2-quart Mason jars in an anaerobic chamber. The jar lids were fitted with a port and 

valve to allow lowering internal pressure to facilitate sealing the jars. A split of each sample was 

dried in the anaerobic chamber at room temperature. The sediments are comprised of medium 

sand to clay size material that ranged from gray to greenish gray color, and were non-cohesive. 

The sand fraction appeared to include dark mineral grains. The samples recovered by washing 

out of bore holes included varying amounts of blue-green blebs of clay which could not be 

removed from the samples. This material was suspected to be from bentonite clay (drilling mud) 

added to facilitate drilling and not from the natural deposit. The sample recovered from the 

corer (3712 BC 580-590) did not contain the clay blebs. The un-dried aquifer materials were 

stored at 4°C in the absence of oxygen in the Mason jar. An oxygen indicator was placed in 

each jar and in an empty jar to verify that oxygen was not leaking into the sample. In addition, 

internal pressure measurements prior to opening jars for subsampling indicated that jars 

remained sealed during storage.

The experimental studies were intended to investigate the ability of the indigenous microbial 

population for biostimulation reduction and were conducted with the 3712 BC 580-590 ISR 

aquifer sediment. Sediments were handled aseptically within the anaerobic chamber. In 

addition all materials were sterilized prior to use, if possible. In particular, storage jars and 

column apparatus were sterilized by autoclaving. Subsampling implements such as spatulas 

were sterilized by wiping with alcohol swabs to minimize contamination of microbial community. 

Subsampling from the jars was conducted in the anaerobic chamber using sterilized apparatus.

3.2.1.2 Groundwater sampling and analysis

Groundwater was sampled from six wells near the drilling site at the Kingsville Dome ISR 

operation, August 5-7, 2009. The wells are screened within reduced (13402, 13406, 13307, 
3-7 

13209) and oxidized (13408, 13310) zones of the roll front deposits. In situ leaching of uranium 

deposit has occurred in this part of the aquifer. Wells were pumped at 75 L/min for a minimum 

of 3 casing volumes. Groundwater samples were collected off of a tee-fitting in the pump 

effluent tubing and filtered through 0.45 micron cartridge filters using a peristaltic pump. Ground 

water dissolved oxygen concentration was measured using Chemetrics dissolved oxygen vacuvial filled with Rhodazine DTMwhich has a minimum detection limit of 3.2 x10-6 M (0.1 mg/L). 

The well sampling pump outflow, split from the TEE fitting, was routed through an inverted 

funnel (~30 mL) that was provided with the vacu-vial ampoules. The ampoule tip was 

submerged in the funnel and broken underwater after 3 minutes of flow through the funnel at 

about 100 mL/min. After the 1 minute color development period, absorbance of the dissolved 

oxygen complex was measured on a Chemetrics V2000 spectrophotometer.

The groundwater samples all had similar major ion chemistry and pH (Table 3.1). Low to 

non-detectable dissolved sulfide and very low dissolved oxygen (<0.3 mg/L; <9 μM) were 

measured in these wells. These low dissolved oxygen concentrations are consistent with the 

relatively high observed Fe(II) concentrations. Total alkalinity ranged from 12 to 15 milliequivalents per liter. Groundwater was near saturation with respect to calcite and gypsum, with 

very high calculated pCO2 (35 to 51%) for the measured pH and alkalinity.

Dissolved iron ranged from 165 to 380 μM (7.9 to 21 mg/L). Field measurements of ferrous 

iron all exceeded the analytical limit of 6 mg/L. Siderite was slightly oversaturated in all samples 

(log SI 0.2 to 0.5). Dissolved uranium, [U], ranged from 20 to 53 μM (4700 to 12,500 μg/L). 

Assuming the dissolved U is U(VI), its speciation is dominated (>97%) by calcium uranyl 

carbonate complexes based on equilibrium speciation modeling of groundwater chemistry using 

PHREEQC and stability constants in Davis and Curtis (2007). Dissolved Mo, likely molybdate, 

ranged from 12 to 48 μM (1200 to 4,600 μg/L). Vanadium was below the ICP-OES method 

detection limit (0.5 μM or 25 μg/L). There is no apparent trend among U, V and Fe concentration 

for these wells.

Groundwater data provided by URI for this well field has similar concentration ranges for Ca, 

Cl, sulfate, alkalinity and U for the most recent samples collected (June 2008) to those shown in 

Table 3.1. The pH values previously reported are 0.7 to 1.1 units higher likely because of CO2

outgassing prior to pH measurement. Groundwater pH data presented here was collected in a 

flow cell that had continuous flow from a diversion of the well pump outlet (at a rate of 100 to 

200 mL/min through the flow cell).
3-8 

Table 3.1 Kingsville Dome groundwater chemistry from wells sampled August 5-6, 2009. Major 

and minor cations determined by ICP-OES of filtered (0.45 μm), acidified samples. Anions 

determined by ion chromatography of filtered samples. The following elements are not reported 

since below method detection limits: Co, Cu, P, F (< 2E-8, 8E-8, 1E-6 M, 5E-7M). “n/m” 

indicates not measured.

Well 13402 13406 13307 13208 13209 13310 units

Field parameters

pH 6.12 6.22 6.13 6.19 6.09 6.10

dissolved O2 <0.1 <0.1 <0.1 <0.1 0.19 0.24 mg/L

H2S n/m n/m n/m n/m 0.17 mg/L

ferrous iron >6 >6 >6 >6 >6 >6 mg/L

Alkalinitya 1.39E-02 1.16E-02 1.20E-02 1.16E-02 1.12E-02 1.24E-02 eq/L

Alkalinityb 1.45E-02 1.19E-02 1.26E-02 1.19E-02 1.20E-02 1.30E-02 eq/L

Ca 1.87E-02 1.78E-02 1.79E-02 1.83E-02 1.88E-02 2.01E-02 mol/L

K 6.36E-04 7.31E-04 5.87E-04 7.77E-04 7.60E-04 7.18E-04 mol/L

Mg 4.53E-03 4.33E-03 3.87E-03 4.87E-03 4.58E-03 4.76E-03 mol/L

Na 2.58E-02 2.75E-02 2.43E-02 2.64E-02 2.77E-02 2.59E-02 mol/L

Si 3.89E-04 4.01E-04 3.57E-04 5.01E-04 4.80E-04 5.48E-04 mol/L

Cl 1.89E-02 2.02E-02 1.82E-02 2.00E-02 2.01E-02 1.88E-02 mol/L

SO4 2.12E-02 2.21E-02 2.10E-02 2.42E-02 2.33E-02 2.26E-02 mol/L

U 5.25E-05 2.93E-05 1.97E-05 2.66E-05 2.86E-05 3.06E-05 mol/L

Fe 2.92E-04 1.65E-04 3.16E-04 1.41E-04 3.83E-04 3.25E-04 mol/L

Mo 1.22E-05 3.45E-05 1.33E-05 4.35E-05 4.84E-05 2.42E-05 mol/L

Ni 2.21E-06 1.70E-06 1.36E-06 1.19E-06 8.52E-07 6.81E-07 mol/L

Sr 2.19E-04 2.02E-04 2.01E-04 1.85E-04 2.08E-04 2.24E-04 mol/L

V <5E-07 <5E-07 <5E-07 <5E-07 <5E-07 <5E-07 mol/L

Zn 2.57E-05 1.93E-05 1.54E-05 1.33E-05 7.95E-06 6.88E-06 mol/L

Ionic Balancec -1.3 -2.2 -2.6 -4.1 -1.6 0.2 %

PHREEQC calculations:

pCO2 0.51 0.35 0.45 0.36 0.46 0.49 atm

log SI calcite -0.01 -0.01 -0.06 -0.05 -0.13 -0.05

log SI siderite 0.49 0.28 0.5 0.17 0.51 0.47

log SI gypsum 0.01 0.01 0.01 0.05 0.04 0.06

a. Measured total alkalinity in samples after return to Menlo Park. Iron oxidation and precipitation occurred 

during this time interval.

b. Measured total alkalinity corrected for proton release resulting from iron oxidation and precipitation by 

assuming 2 equivalents H+ released per mole Fe.

c. Ionic balance uses corrected alkalinity values.
3-9 

3.2.2 AGW Design and Composition

Artificial groundwater (AGW) used for both batch and column experiments was developed to 

simulate the average major ion chemistry of groundwater sampled at the Kingsville Dome ISR 

site (Table 3.1). AGW composition is shown in Table 3.2. The pCO2 was fixed based on 

Kingsville groundwater with dissolved calcium concentration iteratively adjusted in PHREEQC 

simulations to yield the desired pH of 6.3 for the column influent. The resulting calculated 

alkalinity of 5.6E-3 M was about a factor of 2 lower than groundwater from the field site (1.1 to 

1.3E-2 M) which had a calculated pCO2 range of 35 to 51%. The 15% pCO2 was chosen was 

chosen to avoid the need to pressurize the experimental apparatus to work at the higher pCO2. 

Both calcite and gypsum, phases present in the aquifer material, are slightly undersaturated in 

the AGW with log Saturation Index (SI) of -0.23 for both, but were calculated to be at saturation 

in the Kingsville aquifer. For the biostimulation stage of the column experiments and for batch 

experiments, the AGW was amended with dissolved phosphate, ammonia, trace elements, and 

vitamins, as shown in Table 2. Dissolved uranium as U(VI) was added from a 1 mM stock 

solution prepared from dissolution of reagent grade UO3 with sulfuric acid.

The initial batch experiment was conducted at pH 6.9 because of availability of 3.5% pCO2

gas mixture. The AGW used for this initial batch experiment had the total carbonate and calcium 

lowered to 4.62E-3 and 8.32E-3M, respectively, in order to maintain condition of saturation with 

respect to calcite for this CO2 partial pressure. 

AGW for columns was prepared in 3.5 L batches in tared 5-L glass reservoirs (Kontes model 

KC14395 5000) fitted with 3-hole screw caps (Kontes 953930) that were threaded for gas inlet 

and outlet, and liquid outlet tubing fittings. Calcium sulfate was dissolved in 3 L deionized water, 

with appropriate volumes of concentrated salt solutions then added to yield desired final 

concentrations of all major ion components except for NaHCO3 and U(VI). The reservoir was 

then sterilized by autoclave, along with inlet and gas purging tubing assemblies. After 

autoclaving, the reservoir was sealed until cooled. The remaining AGW components and 

amendments were then added through 0.22 μm filters to sterilize them. Reservoir weight was 

recorded after each addition, and after autoclaving. For the lactate column influent reservoirs, 

sodium lactate solution was added from a 1 M stock solution prepared from lactic acid and 

sodium hydroxide. This stock solution was filter sterilized and stored at 4° C in an autoclaved 

serum bottle flushed with nitrogen. After addition of all components, the AGW reservoirs were 

flushed with 15% pCO2 balance N2 gas mixture for one hour at ~1 L/min through a gas 

dispersion stone. The inlet tubing apparatus was installed, and the reservoir flushed for another 

30 minutes before attaching to a column. A continuous flow of the gas mixture was bubbled 

through the reservoir at ~50 ml/min and vented through an airlock for the duration of the 

experiment. 

The hydrogen amended columns used the same AGW composition and preparation method 

but without addition of sodium lactate. Suboxic AGW used to test remobilization of sequestered 

uranium was prepared by the same method but without the additions of dissolved U(VI), 

electron donor, nutrients, vitamins and trace elements. The suboxic AGW reservoir was purged 

with the desired gas mixture (see below).
3-10

Table 3.2 Artificial groundwater composition for column and batch experiments. Lactate, 

nutrients, trace elements and vitamins added only to AGW during biostimulation.

Constituent moles/L

without 

lactate

Ca 2.05E-02

Na 2.96E-02 1.96E-02

Mg 4.87E-03

K 7.77E-04

U 2.00E-05

NH4 1.40E-04 0

Cl 4.52E-02

SO4 1.00E-02

Total CO3 1.13E-02

lactate 1.00E-02 0

PO4 8.00E-07 0

pH 6.23

alkalinity 5.59E-03

Trace elements, vitamins

CoCl2•6H2O 8.0E-07

MnCl2•4H2O 5.1E-07

ZnCl2 5.1E-07

H3BO3 9.7E-08

Na2MoO4•2H2O 1.6E-07

NiCl2•6H2O 1.0E-07

CuCl2•2H2O 1.2E-08

CoCl2•6H2O 8.0E-07

p-amino-benzoic 

acid 1.2E-08

biotin 2.7E-09

folic acid 1.5E-09

pyridoxine•HCl 1.6E-08

thiamine•HCl 5.0E-09

riboflavin 4.4E-09

nicotinic acid 1.4E-08

pantothenic acid 3.5E-09

thioctic acid 8.1E-09

vitamin b12 2.0E-11

3.2.3 Batch Experiments to Determine Electron Donors

Uranium uptake experiments were conducted with the ISR sediment to determine if 

indigenous microbial populations in the sediment enhanced U uptake when amended with 
3-11

different electron donors. An increase in U uptake in the presence of an electron donor 

compared to U uptake in the absence of electron donor was inferred to result from microbial 

reduction. These experiments were conducted in batch mode under anaerobic conditions. The 

results were used to guide the choice of electron donor to use in column experiments. Splits of 

the ISR sediment were transferred into sterilized serum bottles followed by addition of 60 mL of 

AGW to yield a solid to liquid ratio of 50 g/L. 

The initial experiment (Batch 1), started in May 2010, used the pH 6.9 AGW and had an 

imposed pCO2 of 3.5%. Twelve serum bottles with 3 g of sediment and 60 mL AGW were 

sealed in the anaerobic chamber, removed from the chamber, then purged with 3.5% CO2/N2

balance gas stream for 30 minutes to remove hydrogen gas from the bottles and to achieve the 

desired pCO2 and pH. The bottles were then put on an orbital shaker. After 24 hours an initial 

sample was collected through the septum cap with a N2 flushed syringe, and filtered through a 

0.2 μm filter. The filtrate was acidified and processed for dissolved U and ICP analysis. Aliquots 

of 1 mM uranyl sulfate were added to each bottle to attain an initial dissolved U(VI) 

concentration of 20 μM in all bottles except two pairs. One pair of bottles had no added U to test 

for U desorption from the ISR sediment. The other pair had U added to 40 μM. Electron donors 

were added to three pairs of bottles. Hydrogen gas was added to one pair of bottles to attain a 

H2 partial pressure of 0.35 for an equilibrium dissolved H2 of 270 μM. Sodium acetate was 

added to two bottles to yield 10 mM initial concentration. Sodium lactate was added to two 

bottles for initial concentrations of 10 and 20 mM to each of two bottles. No electron donor 

amendments were made to the other six bottles: 2 without added [U], 2 with 20 μM [U], and 2 

with 40 μM [U]. The latter two pairs were used as controls to monitor U uptake in the absence of 

electron donor amendment. All bottles with added [U] were then sampled to establish initial 

actual [U]. Bottles were mixed on an orbital shaker at room temperature and subsampled 

periodically over the next twelve weeks initially at 2 or 3 day intervals with longer intervals based 

on changes in [U]. Bottle weights were recorded before and after each addition and sampling. 

At the end of the experiment, sediments were recovered in the anaerobic chamber, frozen at -

80°C for microbial assay.

A subsequent set of batch experiments was conducted to test the viability of sediments in 

response to hydrogen amendment and test the effect of different dissolved H2 concentrations on 

U uptake and reduction. This experiment was initiated in June after no appreciable decrease in 

[U] was observed in the H2 amended column experiment (see Section on H2 columns below). 

The same sediment as in both the column experiments and in the first batch experiment, and at 

the same solid to AGW ratio (50 g/L) was used. Both the 3.5% and 15% CO2 AGW were used 

to test for any difference resulting from pH 6.9 and 6.3. An initial [U] of 20 μM was used in all 

bottles and was added with the AGW. A range of dissolved H2 also was tested that spanned the 

dissolved H2 imposed on the column reservoir (5%), to the level added to the first batch 

experiment (~40%), and an intermediate level (20%). These H2 partial pressures in the 

headspace resulted in equilibrium dissolved H2 of 40, 312, and 156 μM, respectively. Batch 

bottles were processed and subsampled as described for the first experiment. Sediments were 

not recovered for microbial assay.
3-12

3.2.4 Column Set Up and Operation

Five columns of identical dimensions were packed for this study. Two columns (L1 and L2) 

had lactate added as an electron donor to stimulate indigenous microbes for reduction of 

uranium. One (L1) was subsectioned at the end of the biostimulation stage and sediments 

recovered for geochemical and microbial analyses. The second column (L2) was used to test 

remobilization of sequestered uranium (re-oxidation of U(IV) and desorption or dissolution of 

U(VI)) in response to suboxic conditions to simulate conditions in an aquifer following cessation 

of electron donor addition. This is termed the oxidation stage. A second pair of columns was set 

up to investigate biostimulated uranium reduction by dissolved hydrogen as the electron donor 

and subsequent remobilization under suboxic conditions (H1 and H2). The fifth column (L5) was 

used for a conservative anion breakthrough test (bromide) to determine column hydrodynamic 

parameters prior to biostimulation, and to determine the extent of mobilization of the remaining 

uranium associated with the aquifer sediment following in situ leaching by imposing suboxic 

conditions following the bromide tracer breakthrough and elution. 

3.2.4.1 Columns

Glass 2.2 cm inner diameter columns with adjustable length bed supports were used 

(Millipore VL22X250). Inlet and outlet bed supports were 20 micron pore size. 

3.2.4.2 Plumbing and pumps

Column inlet and outlet tubing was 1/16’’ OD PTFE with 1⁄4”-20 flangeless fittings to connect 

to 4-way valves at about 12” from column inlet and outlet fittings. A pressure gauge was 

installed at a TEE fitting upstream of the inlet 4-way valve and downstream of the pump. All 

tubing from the reservoir cap to effluent collection outlet was encased within 1⁄4” ID 1/8” wall 

Tygon R3607 tubing to jacket the influent and effluent tubing. The gas mixture was continuously 

flowed thoughout the jacketing to minimize gas exchange through influent and effluent tubing 

and oxygen contamination. Tygon tubing was attached to nylon TEE connectors at each fitting 

and at the column inlet and outlet with a length of Tygon connecting adjacent TEE connectors to 

provide a continuous purge of the entire length of influent and effluent tubing. The jacketing 

purge gas was flowed at 50 cc/min in the same direction as the AGW flow starting at the feed 

line of the reservoir to the column effluent outlet, bypassing the peristaltic pump, the pressure 

gauge, valves and the columns (see Figure 3.4).

The influent AGW was pumped from the reservoir through the columns using an ISMATEC 

8-roller multi-channel pump fitted with 1.1 mm ID and later 0.76 mm ID Tygon R3606 double 

stop tubing that was attached to jacketed reservoir and inlet tubing with nylon hose barbs. All 

tubing and fittings were autoclaved before use.
3-13

A B C

D

E

F

Figure 3.4 Column experimental apparatus illustrating (A) column and effluent collection, 

(B) effluent tubing with gas purged jacketing to minimize oxygen invasion and effluent 

collection bottle, (C) influent reservoir with gas mixture purge and outlet lines, (D) influent 

lines with gas jacketing, (E) peristaltic pump in purged housing. F) Column L2 after 46 

days of biostimulation showing dark banding forming in sediment (arrow) and Sterivex 

filter on outlet for microbial assay test in gas jacketed housing upstream of effluent 

collection bottle.
3-14

3.2.4.3 Gas phase and oxygen control

Commercially-prepared, high-purity gas mixtures of 15% CO2, 85% N2 were used for the 

biostimulated reduction stage of the lactate columns and 15% CO2, 5% H2, 85% N2 was used 

for the hydrogen amended column experiments. An oxygen scrubber was fitted in line to 

remove residual oxygen in the commercial mixture. Reservoirs were continuously purged with 

the gas mixture at about 25 cc/min, after the initial flush, and vented through airlocks. Jacketed 

tubing (described above) also was continuously purged with this gas mixture. For the suboxic 

elution of the lactate columns, the gas stream was delivered using a 3-channel gas flow 

controller to dilute a 20% O2, 15% CO2, 65% N2 premixed cylinder with N2 and CO2 to attain the 

desired oxygen content while maintaining 15% CO2. The pO2 of the mixture was set to attain 

the desired dissolved oxygen concentration. The mass flow controller was calibrated using a 

TCD detector gas chromatograph. Because of gas exchange across pump tubing walls resulting 

from pressure from pump rollers, the peristaltic pump was housed within an acrylic box that was 

continuously purged with the gas mixture for the specific experiment. This housing was 

constructed for the hydrogen amended columns and later used for the suboxic elution of lactate 

column experiment. Significant loss of hydrogen and influx of oxygen was discovered during the 

course of the hydrogen biostimulation and the suboxic elution of lactate experiments, 

respectively, as discussed below. The purged pump housing eliminated the problem with gas 

exchange across pump tubing. 

3.2.4.4 Column packing

The sediments from core KVD 3712 BC 580-5590 were used for column biostimulation 

experiments. Columns were slurry packed with wet aquifer sediment in an anaerobic glove 

chamber. Sediments were mixed with AGW to a consistency similar to pancake batter by adding 

about 20 mL of AGW to 100 grams of damp aquifer sediment and mixing thoroughly. Starting 

and ending weights of the sediment slurry container were recorded during column packing. The 

sediment slurry was subsampled to tared glass vials to determine water content upon drying. 

Each column with its inlet and outlet fittings and tubing including 4-way valves at each end was 

weighed empty. Dead volumes of tubing, frit and fittings were determined by weighing the 

column assembly full of deionized water with the inlet and outlet bed supports adjusted inwards 

until they met and subtracting the empty weight. 

An empty column with the inlet bed support and fittings attached was mounted vertically 

with the inlet end down. AGW was added through the inlet valve with a syringe until a few 

millimeters above the inlet bed support frit. The sediment/AGW mixture was added with a 

spatula to the column in increments of about 0.5 cm. A rubber policeman was used to move any 

sediment that adhered to column walls to the sediment bed. The column was tapped and 

sediment was allowed to settle for about 5 minutes before adding the next increment. AGW 

was added as needed to maintain a lens of water (1-2 mm) above the sediment. This process 

was repeated until the sediment was at a height of about 10.5 cm above the inlet frit. The 

sediment in the packed column was allowed to settle overnight. On the following day, any 

overlying water (~5 mL) was removed from the top of the sediment bed with an automatic 

pipette and retained in a tared vial to determine the mass of water and sediment removed. 

Small amounts of fines were inadvertently removed by this process. The walls of the column 
3-15

were cleaned with damp tissue. The outlet bed support and fittings were installed with frit 

adjusted to the level of the top of the sediment. Additional AGW was pushed through the 

column by syringe. The packed column was allowed to stand for an additional 24 hours. The 

outlet frit was then adjusted downward to the sediment surface as needed to expel excess 

overlying water through the outlet tubing. Inlet and outlet valves were then closed and the 

column was removed from the anaerobic chamber and weighed. Columns are mounted 

vertically with the outlet end up, and the influent and effluent plumbing attached to the 

appropriate 4-way valves. Flow of the gas mixture through the inlet and outlet tubing jackets 

was initiated, and flow of influent AGW containing dissolved U(VI), electron donor and other 

amendments through column commenced.

3.2.4.5 Column parameters; pore volume, solid density

Measured dimensions and weights of columns used for lactate experiments are shown in 

Table 3.3. Column pore volume and porosity were determined from total sediment weight used 

in packing the columns, total weight of column minus empty weight, and the dimensional volume 

minus volume of bed support fittings, inlet and outlet tubing. Pore volume and total dry sediment 

mass were similar in the two columns, which were intended to be identical.

Table 3.3 Column parameters: dimensions and weights. 

L1 L2 L5 Parameter

2.22 2.22 2.22 Column diameter, cm

10.5 10.1 10.1 Column length of sediment, cm

78.50 78.13 79.04 Total sediment weight, wet, in column

0.28 0.28 0.27 Water weight of sediment, from split during packing

56.86 56.59 58.01 Calculated dry sediment in column, g

17.8 16.2 17.9 Calculated water component of wet sediment in column, g

1.58 1.58 1.58 Dead volume of inlet and outlet tubing ,measured; cm3

17.8 16.2 17.9 Pore volume of packed column, cm3

40.6 39.1 39.1 Total column bed volume (from measured dimensions) , cm3

0.44 0.41 0.46 Column porosity, calculated

2.49 2.47 2.74 Sediment density, calculated, g/cm3

3.2.5 Column Effluent

3.2.5.1 Sampling

Samples of column effluent were collected on a continuous basis throughout the duration of 

the reduction of the column experiments into tared autoclaved serum bottles that were flushed 

with the 15% CO2 balance N2 gas mixture. The outlet tubing was fitted with a sterile 0.22-μm 
3-16

pore size 25-mm diameter sterile cartridge filter with an attached hypodermic needle that was 

inserted into 1-cm thick butyl rubber serum bottle stopper. A second needle connected to an 

airlock was inserted that allowed the headspace of the bottle to vent during sample collection

(Figure 3.4A). This collection method was intended to maintain anaerobic conditions in the 

effluent sample bottle during collection and subsequent storage. Sample bottles were changed 

every 24 to 60 hours. Filters were replaced as needed with about three dead volumes allowed 

to flow through filter and needle before attaching to the collection bottle. Sample volume was 

determined by the weight of effluent collected, and the flow rate from volume divided by the 

duration of sample collection. Every other sample was acidified to 1% (v/v) with nitric acid for 

dissolved U and dissolved cation analyses. Samples were acidified at least 24 hours prior to 

subsampling analysis, and may have been stored at room temperature for up to two weeks 

before subsampling. The other samples were stored under refrigeration and were not acidified. 

These unacidified samples were used for analysis of dissolved sulfate, headspace and total 

dissolved carbonate, and, lactate and acetate (columns L1 and L2). Separate samples for 

dissolved sulfide analysis were collected periodically into 5-mL glass vacuvials previously 

flushed with the gas mixture and containing 1 mL 10% (wt/v) zinc acetate to preserve sulfide by 

precipitating as zinc sulfide. 

Following the biostimulation reduction stage, a bromide tracer in AGW without lactate or U 

was passed through column L2 and then eluted with AGW with no U or lactate (see below). 

Samples were collected during the tracer input and elution using a fraction collector. 

Subsequently, AGW with no U, lactate, or added dissolved oxygen was passed through column 

L2 for 32 days with effluent collected in serum bottles, as described above. Effluent samples 

during the addition of dissolved oxygen to the influent (oxidation stage) of column L2, and during 

all of the column L5 sub-oxic elution were collected in plastic scintillation vials using a fraction 

collector programmed at the desired sampling interval. Sample vials were tared before use, 

capped and weighed after sample collection. Correction was made for evaporative loss during 

and after sample collection prior to capping based on evaporative loss from control vials 

containing similar volume of AGW. Samples for U and dissolved cations were acidified to 1% 

(v/v) with nitric acid. 

Influent reservoirs were sampled through the 4-way valve between the pump and the 

column when each new reservoir was installed.

Dissolved oxygen was measured in both the column influent (pump outflow) and column 

effluent. Samples were collected in 7-cc serum bottles through syringe needles extending to 

near the bottom of the bottle and the exit needle inserted just below the stopper. These oxygen 

bottles were overfilled two to three bottle volumes and had negligible gas headspace volume.

3.2.5.2 Bromide tracer

Measurement of the breakthrough and rise to plateau of a conservative anion (bromide) and 

subsequent elution was conducted to determine column hydrodynamic parameters prior to 

biostimulation (column L5). A bromide tracer was also passed through column L2 after 

biostimulation, but prior to suboxic elution, to determine if changes in pore volume and 

dispersion occurred during biostimulation. A reservoir containing bromide (360 mg/L) in AGW 

(no U, lactate, vitamins or nutrients) equilibrated with 15% CO2/N2 gas mixture was flowed for 6 
3-17

pore volumes through column L5 and for 10 pore volumes through the column L2. Continuous 

samples of effluent were collected using a fraction collector at 4 hour intervals for 3 days, 

followed by 6 hour intervals for the duration of the bromide pulse. Following the bromide pulse, 

a reservoir with AGW with no bromide, U, or lactate was installed and equilibrated with 15% 

CO2/N2mixture. Bromide elution samples were collected at the same frequency as bromide 

pulse sampling. After 8 4-hour samples, the effluent collection time was increased to 6 hours 

with effluent samples collected over the next 4 days to capture the tail of eluted bromide. Splits 

of every fourth or fifth sample were acidified for [U] analysis. Oxygen was added to the gas 

mixture of column L2 after 32 days of elution with AGW with the no oxygen gas mixture (see 

below). Bromide sampling of column L5 ended at 11 pore volumes (PV) (day 13). Elution of 

column L5 with suboxic AGW commenced by including oxygen in the reservoir purge gas 

mixture (0.5% O2/15%CO2/N2) for a 0.4 mg/L (12 μM) dissolved oxygen.

3.2.5.3 Analyses

Dissolved U in effluent and batch samples was measured using a kinetic phosphorescence 

analyzer (Chemchek Instruments Model KPA1), following pretreatment of an aliquot of acidified 

sample. The pretreatment consisted of drying on a hotplate in a glass scintillation vial, 

reconstituting in 1 mL of concentrated nitric acid, adding 7-10 drops of 30% H2O2, and heating 

to dryness. The dried sample was reconstituted in 0.1N HNO3 and sonicated prior to KPA 

analysis. The pretreatment removes constituents (e.g. chloride, lactate) that quench uranium 

phosphorescence. KPA measures only U(VI), but in the unlikely event that any dissolved or 

colloidal U(IV) was present in the effluent samples it was likely oxidized to U(VI) during the 

pretreatment process and also measured as [U]. 

Dissolved cation concentrations (Fe, Ca, K, Mg, Mn, Na, etc) in effluent samples were 

measured by ICP-OES. The ICP-OES also measures total dissolved sulfur which includes both 

sulfate and sulfide species. Dissolved sulfide concentrations were measured in preserved 

samples by the colorimetric method described by Cline (1969). Dissolved sulfate in the unacidified samples was measured by ion chromatography, while pH and alkalinity were 

calculated from the pCO2 and total dissolved carbonate measured in the un-acidified samples. 

This method was used because lactate, acetate, and sulfide contribute to acid neutralization 

during Gran titration. Headspace CO2 of un-acidified samples was measured on TCD detector 

gas chromatograph prior to other subsampling from these bottles. Subsequently a 5-mL aliquot 

of the effluent sample was transferred by syringe to a N2 flushed serum bottle and acidified. The 

CO2 evolved from dissolved carbonate was measured by GC. The alkalinity and pH measured 

in influent AGW (without lactate) collected by the same method but directly from the pump outlet 

gave comparable values to pH measured by electrode and to alkalinity determined by Gran 

titration on separate aliquots of this AGW. Lactate and acetate in the un-acidified samples were 

measured by HPLC with a UV detector. 

Dissolved oxygen concentrations were measured in the 7 cc serum bottles in the anaerobic 

glove chamber colorimetrically using Chemette dissolved oxygen ampoules and a Chemetrics 

V2000 spectrophotometer. The ampoule tip was inserted into silicone tubing that was 

connected to a Luerlok fitting and a needle inserted to near the bottom of the bottle. A vent 

needle was inserted just through the stopper. On breaking the ampoule tip, sample was drawn 
3-18

into the ampoule in about 5 seconds with an equal volume replaced by the anaerobic chamber 

atmosphere through the vent needle. After the 1 minute color development period, absorbance 

of the dissolved oxygen complex was measured on the spectrophotometer.

3.2.6 Solid Phase Sampling

3.2.6.1 Column Sectioning and Sample Preservation

Sediments were recovered from column L1 at the end of the reduction stage and from 

column L2 and L5 at the end of the oxidation stage for chemical analyses and microbial assay 

(column L1 only). Columns were dismantled and the sediment subsampled in the anaerobic 

glove chamber. While maintaining the column vertically with the outlet end up, the outlet fittings 

up to the bed support were removed and a section of PVC pipe inserted. The column was then 

inverted and inlet fittings removed. The sediment was extruded out of the column into ~2-cm 

vertical sections with push up distance and length of sediment bed measured before and after 

each subsection was extruded. Each subsection was weighed and homogenized by mixing with 

a sterilized Teflon spatula. A split of sediment was taken to determine water content upon 

drying. Each column subsection was about 11 to 12 grams of dry sediment. For column L1, a 

second split of about 1 g was transferred to a whirlpak bag for microbial assay. The remaining 

sediment was split between a 2-mL micro-centrifuge tube for gamma spectrometry, and a glass 

scintillation vial. These containers were double bagged in Mylar pouches containing oxygen 

scrubbers. The bags were heat sealed and stored at -80° C until analyzed.

3.2.6.2 Total U screening by gamma spectrometry

Total uranium content of the column subsections and the KVD 3712 BC 580-590 sediment 

prior to biostimulation (termed pre-column sediment, here after) was determined by gamma 

spectrometry following the method outlined in Fuller et al (1999). The 63 KeV gamma emission 

of the 238U daughter 234Th was measured on a high resolution germanium detector gamma 

spectrometer. This method assumes that the 234Th daughter is in secular equilibrium. Because 

the L1 samples measured immediately at the end of the biostimulation reduction likely were not 

in secular equilibrium, the total uranium of these samples measured by gamma spectrometry is 

considered a lower limit. These total uranium values were used primarily to determine which 

samples to use for X-ray adsorption spectroscopic measurements and as a guide for dilutions 

for chemical extraction analyses. Total uranium concentrations determined by hot nitric acid 

extraction (see below) were used for mass balance calculations and for comparison to loading 

calculated by integrating the difference between influent and effluent dissolved U.

3.2.6.3 Chemical analyses

Five gram splits of recovered sediment from columns were dried at room temperature in the 

anaerobic chamber for chemical analyses. The dried sediment was homogenized and a 1-

gram split was ground with an agate mortar and pestle for nitric acid extraction and total S 

measurement. The unground remainder was used for partial chemical extraction of ferric and 

ferrous iron, acid volatile sulfide, nitric acid and HF total dissolution, and total carbonate 
3-19

analysis. A split of the pre-column sediment was also analyzed by the techniques outlined below 

to compare with this sediment following biostimulation reduction and suboxic elution. 

3.2.6.3.1 Extractions

Hot nitric acid and peroxide extraction was used to determine total uranium and other 

constituents such as calcium and iron. This method is based on EPA method 3050 and 

dissolves most sedimentary phases except for silicates. Briefly, 0.1 gram of the ground 

sediment was placed in a tared glass vial and treated with ~2 mL 0.5 N HNO3 added drop wise 

to minimize samples loss as aerosol during effervescence of carbonates. The aquifer sediment 

contained about 20% (w/w) carbonate as CaCO3. The pre-treated sample was dried at 90°C on 

a hot plate and cooled. 3 ml of concentrated HNO3 was added and a glass bulb placed on top of 

the vial. The sample was refluxed for 24 hours at 90°C. After cooling, 0.5 mL of 30% H2O2 was 

added and allowed to stand for 30 minutes. The reflux bulb was removed and the sample 

evaporated to dryness at 90°C. The dry weight was recorded and 10 mL of 0.1N HNO3 added to 

reconstitute the sample. The sample was sonicated for 30 minutes and allowed to settle 

overnight. An aliquot of supernatant was removed with a syringe, passed through a 0.22 μm 

filter, and diluted appropriately for analysis by KPA for U and ICP. The U concentration 

determined using this hot nitric acid extraction and the gamma spectrometry total U 

measurement generally agreed within 10%. 

Total dissolution via hydrofluoric acid digest was also performed to provide a measure of 

iron content of silicate phases not leached by the hot HNO3 digest. Briefly 0.1 g of ground 

sediment was placed in a 10 mL Teflon reaction vial. Samples were pre-treated to remove 

carbonates as described above, then dried. Samples were then digested in a mixture of 2.5 mL 

concentrated HF and 0.5 mL concentrated HNO3 at 90°C for 4 days. After drying, samples were 

reconstituted in 5 mL of 0.5M HCl, and then processed for ICP and KPA analyses as described 

above. HCl is required to redissolve the large amount of calcium fluoride formed during the 

reaction of Ca from carbonates with HF. 

Extractable ferrous (Fe(II)) and reducible iron were determined by 0.5 N HCl and 0.5 N 

HCl/0.25M hydroxylamine hydrochloride (HH) extractions of un-dried, unground sediments, 

respectively. The difference between these two extractions has been defined operationally as 

the poorly crystalline ferric iron available for microbial reduction (Lovley and Phillips, 1987). The 

extractions were conducted in the anaerobic chamber on anaerobically preserved sediments. A 

separate split of each sample was taken for drying to determine water content to allow reporting 

concentrations as per gram of dry sediment. Briefly, 0.5 g of sediment was transferred to a tared 

polycarbonate Oak Ridge centrifuge tube and 30 mL of either 0.5 N HCl and 0.5N HCl/0.25 M 

hydroxylamine hydrochloride was added. Each sample was extracted in duplicate using each 

solution. Tubes were mixed on an end over end shaker (12 RPM), and 3 mL of extraction 

solution recovered by filtering through 0.22 μm syringe filter. Extractions with 0.5 N HCl were 

sampled at 1 and 24 hours. HH extractions were sampled after 1 hour and 24 hours. Tubes 

were weighed after each subsample to account for the volume removed. Fe(II) concentrations 

in the HCl extraction solutions were determined by the ferrozine colorimetric method (Stookey, 

1970). The total iron concentration in the 0.5 N HCl extraction solutions was determined by 

adding hydroxylamine hydrochloride to a separate aliquot of the filtered extraction solution to 
3-20

reduced dissolved Fe(III), and analyzed by the ferrozine method. Total Fe extracted by HH was 

determined directly by ferrozine since extraction reduces Fe(III).

3.2.6.3.2 Carbonate, total sulfur, acid volatile sulfur

Total carbonate content of the pre- and post-column sediments was determined by 

measuring the CO2 evolved upon acidification. Briefly, 0.5 to 1.0 g splits of dried sediment were 

placed into 120 cc serum bottles which were sealed and flushed with N2. Five mL of 2N HNO3

were injected into the bottle after removing an equal volume of headspace. After mixing on an 

orbital shaker, a 60-cc syringe fitted with a 3-way valve and a needle was inserted into the 

stopper. The excess gas pressure was allowed to displace the plunger upwards and volume of 

displacement recorded. After closing the valve and removing the syringe from the bottle, the 

syringe was connected to the injection loop of a TCD detector GC to measure CO2 content of 

the gas phase. The percent carbonate in the solid was calculated by multiplying the gas phase 

CO2 concentration by the sum of the bottle volume and syringe displacement and dividing by the 

initial sample weight. 

Total sulfur content and S isotope ratio of the dried, ground column sediment and precolumn material were measured using a Carlo Erba elemental analyzer (EA) coupled to a 

Micromass Isoprime mass spectrometer. This method provides a measure of all forms of S 

present in the sample. An increase in the % total S would reflect reduction of sulfate in the AGW 

during biostimulation and precipitation of reduced sulfur species, such as FeS. The S isotopic 

ratio, expressed as δ34S, can be used as an indicator of fractionation of S isotopes during 

microbial reduction of S (Habicht and Canfield, 1997), resulting in a lighter isotopic ratio (lower 

δ34S) of the reduced S. This fractionation results from the kinetic isotope effect in which the 

reaction rates for heavier isotopes (e.g 34S) are slower resulting in a depletion of the heavier 

isotope in the product relative to the reactants. Briefly, 3 mg of the homogenized ground 

sediment sample was weighed on a microbalance into tin capsules with approximately 2 mg of 

V2O5 added. The samples are combusted at 10000C. The EA traps water and separates the 

resulting CO2 and N2 from the SO2 which is measured using an IR detector and then sent to the 

mass spectrometer. The N2 and CO2 are vented. The samples are run in batches of 30 at 10 

sample intervals. Each interval is surrounded by four reference standards (NBS SRM bovine 

liver) run in a range of sizes to capture the range of S in the samples. A blank is included in the 

run. The standards are calibrated against Canyon Diablo Troilite (CDT) forδ34S. The 34S isotope 

composition is expressed as a difference in the ratio relative to the CDT in parts per thousand or 

per mil (‰).The δ34S and %S values are corrected for instrumental drift and size linearity. δ34S 

values are also corrected for oxygen contribution. One standard deviation of the standards 

throughout a run is ± 0.5 ‰ or better for δ34S, and 0.01 %S. Duplicates measured every fifth 

sample agreed within 0.05% S and within 0.5 per mil δ34S . 

Acid volatile sulfur (AVS) content was measured on sediments recovered from the lactate 

columns (L1 and L2), and the pre-column sediment using a modified version of the diffusion 

method of Hsieh and Yang (1989). AVS is a measure of sulfide (S-2) species in the sample 

such as FeS that are liberated upon acidification forming H2S gas. The method used does not 

reduce more oxidized forms of S such as elemental sulfur or S-1 in FeS2. In the anaerobic 

chamber, dried column sediment samples (0.3 g) are placed in 100 mL serum bottles along with 
3-21

1 mL of 1 M ascorbic acid to eliminate oxidation of AVS by ferric minerals (Hsieh et al, 2002). A 

10 x 75 mm test tube containing 3 mL of alkaline zinc acetate (3% w/v zinc acetate in 2N NaOH) 

is placed in the serum bottle, with opening upwards and resting on the shoulder of the bottle. 

The alkaline zinc acetate solution traps the H2S gas evolved on acidification forming a zinc 

sulfide precipitate. A 1-cm magnetic spin bar is placed in the bottom of the bottle. The serum 

bottles are then sealed and 10 mL of 6N HCl is injected through the serum bottle stopper with a 

syringe directly onto the sediment sample carefully avoiding the contact with the zinc acetate 

solution. The bottles are removed from the anaerobic chamber and placed on a stir plate with 

gentle stirring. After 30 hours, the zinc acetate tubes are recovered. After weighing each tube, 

an additional 0.5 mL of 2N NaOH is added, and the tubes are stoppered, then sonicated for 30 

minutes to disperse the ZnS precipitate. Weights are recorded at each step to determine exact 

volumes. The resulting suspension of zinc sulfide is subsampled and sulfide is determined 

colorimetrically by the Cline method (see above). The method was tested using a reference 

synthetic solid (70% FeS; 30%FeS2). The measured AVS within 24 hours was equivalent to the 

S in the FeS component. No additional S was recovered after longer equilibration times 

consistent with previous studies that acid without added reductants only volatilizes the S-2, and 

not higher oxidation states such as S- in FeS2, or elemental S. Replicates of this reference 

sample agreed to within ± 5%, and duplicates of column samples agreed within ±10%. The 

effect of CO2 evolved from the carbonates present in the ISR samples during acidification on 

AVS recovery by alkaline zinc acetate was found to be negligible. Other treatments such as 

acidic Cr(II) to reduce higher oxidation states of S were not attempted. Instead, the difference 

between total sulfur increase (column minus pre-column total S) and the AVS is used as a 

measure of higher oxidation states of S precipitated during the biostimulated reduction stage of 

the column experiments. 

3.2.7 Microbial Assay

Microbial characterization of sediments and effluent filters were conducted by the USGS 

microbiology laboratory in Reston, VA. The general approach taken was to extract DNA, and 

perform bacterial Terminal Restriction Fragment Length Polymorphism (TRFLP) fingerprinting 

and quantitative Polymerase Chain Reaction (qPCR) analyses of Geobacter and sulfate 

reducing bacteria (SRB) on all samples. In addition selected samples were further characterized 

by cloning Geobacter 16S rRNA gene, and/or dsrB gene coding for the dissimilatory sulfite 

reductase involved in sulfate reduction. The intent was to 1) gain an understanding of microbial 

dynamics in these experiments as biostimulation of U(VI) reduction occurred, as well as during 

potential remobilization of uranium, 2) assess where in the column the microbial abundances 

were highest and relate that to chemistry and solid phase analyses, and 3) determine potentially 

important Geobacter sp. and sulfate reducing organisms involved in these processes.

The sample sets included sediments recovered from the Kingsville Dome ISR site collected 

in August, 2009, after in situ leaching had been stopped for about two years. Sediment samples 

recovered at the end of the first batch experiments with the 3712B-C 580-590 ISR core material 

which were amended with acetate, lactate, or H2. Sediments recovered from column L1 at the 

end of the biostimulation stage also were assayed, as well as filters of effluent collected from 

column L2 during lactate addition. The first filter was started at the end of the period of high 
3-22

effluent ferrous iron concentration and continuing for 20 days (19.5 PV). The second filter was 

collected over the next 14 days during which time effluent sulfide began to increase. 

3.2.7.1 Batch and column sediments

Sediment was subsampled from batch experiment 1 bottles at the end of the 60 day 

experiment. After settling, the overlying water was removed in the anaerobic chamber. About 

one gram of wet sediment was transferred to a whirl pak bag using a sterilized spatula. 

Samples were sealed in Mylar bags and frozen at -80°C. One gram splits of column L1 

sediments were sampled immediately upon sectioning of the column in the anaerobic chamber. 

The subsamples were processed and preserved as described for batch samples. 

3.2.7.2 Effluent filters

Millipore Sterivex 0.2 μm filter capsules were installed in the column L2 effluent line within a 

gas purged holder to limit exposure to oxygen. The effluent was collected as described above, 

downstream of the Sterivex filter. Following the collection period, the filter capsules ports were 

sealed and the cartridge was stored at -80°C. Filter capsules and sediment samples were 

shipped on dry ice to the microbiology lab.

3.2.7.3 Measurement methods

DNA Extraction. The 10-20g each of frozen aquifer sediments samples and the 0.5-1g 

frozen sediment samples from the batch and column experiments were thawed and then 

extracted using the MoBio ultraclean soil DNA megaprep kit according to manufacturer’s 

instructions (MoBio, Inc., Carlsbad, CA). Sterivex capsule filters of column L2 effluent were 

thawed and extracted as previously described using the Qiagen (Gentra) puregene kit with 

slight modifications. (Qiagen, Inc., Valenica, CA) (Ward et al. 2007). Briefly, 0.9mL of lysis buffer 

with 4.5μL proteinase K was added to the filters and incubated with gentle rotation for 10 

minutes at 80°C. Volumes of solutions used in subsequent protein precipitation and DNA 

precipitation steps were scaled up accordingly (3X) to reflect the 3X volume increase in lysis 

step compared to manufacturer’s instructions (0.9mL instead of 0.3mL).

Bacterial 16S rRNA Gene PCR and TRFLP Fingerprinting. The polymerase chain 

reaction (PCR) technique was used to amplify bacterial DNA from samples, targeting the 

evolutionally conserved ribosomal RNA gene, 16S rRNA. This technique allows for the 

amplification of a specific segment of DNA of interest from the bulk DNA. Enough of the 

amplification product from PCR (amplicon) is produced such that it can then be detected and 

characterized further (cloning and sequencing or fingerprinting depending on the question being 

asked). PCR was performed as previously described using the 16S rRNA gene primers 46fFAM and 519r (Jones et al 2006, Lane 1991, Brunk et al 1996). A DNA fingerprinting method, 

the terminal restriction fragment length polymorphism (TRFLP) technique was also used to look 

at bacterial community differences between samples. This technique involves the enzymatic 

digestion of the PCR product, which has a fluorescent tag. The enzyme has a specific DNA 

recognition sequence and only cuts the DNA where that sequence exists. Therefore, depending 

on the PCR product’s DNA sequences, enzyme digested products of varying size will be 
3-23

produced. The fragments from the end which contains the fluorescent tag can then be 

electrophoresed (size separated) and the fluorescence detected, generating a fragment size 

profile which is unique to that population of organisms with those sequences. This fingerprint 

can be compared to other samples to evaluate similarities and differences between the sample 

communities or within a community over time. TRFLP was performed as previously described 

(Jones et al 2006). Briefly, PCR amplicons were digested with Mn(II), precipitated, and 

electrophoresed on an ABI 310 genetic analyzer to generate fingerprint profiles. Fingerprint 

data were binned using the software R, and imported into the SAS based JMP8 statistical 

package (SAS, Cary, NC) for clustering analysis.

Real time quantitative PCR (qPCR) for Geobacter and SRB. qPCR was performed on 

all samples for Geobacter using the 16S rRNA gene primers Geo494f and Geo825r (Anderson 

et al. 1998, Holmes et al 2002), and for SRB using the functional gene dsrB primers drp-2060f 

and dsr4r (Wagner et al 1998, Geets et al. 2006). The procedure including cycling 

temperatures and times are as previously published (Wilson et al. 2010). 

Clone Library Construction and Sequencing. Geobacter and SRB clone libraries were 

constructed for selected samples. PCR was performed using the Geobacter specific and SRB 

specific primer sets above. PCR conditions; Geobacter 30 cycles of 94°C (30 s), 53° C (30 s), 

and 72° C (30s) followed by a 7 minute 72°C extension incubation. SRB PCR was also 30 cycle 

and conditions were identical except for the annealing temperature was 56°C rather than 53°C. 

Amplicons were purified by wizard prep (Promega, Madison, WI.), and cloned into TA vector 

according to manufacturer’s instructions (Invitrogen, Carlsbad, CA). Single clones were picked 

and analyzed for insertion by PCR with M13f and M13r primers. PCR products were sequenced 

by single pass PCR sequencing (Beckman Coulter Genomics, Danvers, MA.).

Sequencing Analysis. Plasmid vectors were trimmed using the awk program vbgone 

(Varnum Engineering, Seattle, WA.) Sequences were oriented and aligned in Macvector 12 

using ClustalW (MacVector, Cary, NC). Phylogenetic trees were built in MacVector12 using the 

Neighbor Joining method and TamuraNei distance with bootstrapping (1000 replicates) or 

BESTTREE. Representative sequences from each phylotype were checked for highest similarity 

to sequences in Genbank database using the Basic local alignment tool (Blast).

3.2.8 Spectroscopic Measurements

Sediments recovered from sub-sectioning columns L1, L2 and L5 were used for X-ray 

absorption spectroscopic (XAS) measurement to determine oxidation state and the local 

molecular structure of sequestered uranium. The pre-column material was included for 

comparison.

3.2.8.1 Bulk XAS

X-ray absorption near edge spectroscopy (XANES) and Extended X-ray Absorption Fine 

Structure (EXAFS) measurements were made on anaerobically stored ISR sediments from 

columns and pre-column material at Stanford Synchrotron Light Source (SSRL) beam lines (BL) 

4-1 and 11-2. Analysis of XANES spectra is used to provide a quantitative measure of the 
3-24

different oxidation states of an element in a sample. EXAFS spectra are used to derive the local 

atomic structure surrounding the element of interest in the sample. Qualitatively, EXAFS spectra 

can be compared to reference or model compound spectra to identify the presence of specific 

forms in this case, of uranium. Bulk XAS measurements provide information on the entire mass 

of the element of interest in the portion of sample illuminated by the X-ray beam. Beam size 

was typically 1 mm high by 8 mm wide with the sample placed at a 45° angle to the incoming Xray beam, which resulted in an illumination area of 1 x 11.3 mm for the 1.2 mm thick sample.

In the anaerobic chamber, un-dried samples were ground with agate mortar and pestle, and 

packed into 1/32” thick polycarbonate holders with 10 mil Kapton tape windows. Uranium L-II 

and L-III edge fluorescence spectra were collected in an anaerobic stage to eliminate exposure 

to oxygen during sample collection. 

For XANES, a minimum of three replicate scans of fluorescence spectra of each sample 

was collected across either the U-LII or U-LIII edge using either a 30-element (BL 11-2) or 13-

element (BL 4-1) germanium array detector. Internal calibration of the monochronomator was 

made using either a Y or Mo foil placed on a second ion chamber “down-stream” of the sample 

stage, with foil edges collected at the start of each scan. The sample spectra were deadtime 

corrected and adjusted for drift of the monochronomator prior to averaging. The averaged 

spectra were background subtracted and normalized using SIXPACK software (Webb, 2005). 

The resulting corrected spectra were fit to one or more reference U(IV) and U(VI) model spectra 

using the least squares fitting module in SIXPACK. Spectra of crystalline uraninite, andersonite, 

phosphuranylite and U(VI) sorbed to ferrihyrdrite provided by John Bargar, Stanford Synchrtron 

Radiation Lightsource, were used for model compounds. The model spectra were collected at 

other beam time sessions but also calibrated with internal reference foil for either the U-LII or ULIII edge. The least squares linear combination fitting (LC) procedure yields fractional 

components of U(IV) and U(VI) in the sample. The best fit was based on the lowest residual chi 

squared value. That is, the components those yielded the minimum difference between the 

sample spectra and the fit of the components.

Bulk EXAFS spectra were collected on a limited number of samples because of availability 

of beam time. Up to 12 replicate scans were collected. Spectra were deadtime corrected and 

averaged as described above. The background corrected and extracted EXAFS signal was 

converted to frequency (k) space, weighted by k3

, and Fourier transformed .The low uranium

concentration of the column samples limited the usable data collection range to about k of 9 at 

best. The k3 weighted spectra and its Fourier transform were used to qualitatively compare to 

model compounds. Linear combination fitting and shell by shell fitting of column samples was 

not attempted because of the significant contribution of initial (or background) uranium present 

in the pre-column sediment and the limitations of spectral data quality for lower concentration 

samples. 

3.2.8.2 Microfocused synchrotron XRF and XAS

Thin sections of column sediment were prepared for imaging by microfocused X-ray 

fluorescence (XRF) and for discrete point XANES at the SSRL BL 2-3 and BL 10-2 X-ray 

microprobes, which provide spatial resolution of up to 1 and 20 um, respectively. The 
3-25

microfocused beam XRF (μXRF) provides elemental distributions or maps of elements in a thin 

section. The mapping can be conducted at several energies across the critical X-ray absorption 

edge of the element (e.g. U) in question to provide maps of oxidation state for the area of 

interest (Mayhew et al, 2011). 

Petrographic thin sections of sediment recovered from columns L1 and L2, and the precolumn sediment were prepared by embedding the sediment in epoxy resin, then cut and 

polished to 30-um thickness, and mounted on a quartz slide. Briefly, about 3 grams of the 

anaerobically stored sediment was dried by spreading into a thin layer on a 6”-diameter paper 

filter in the anaerobic chamber. After drying, the sediment was transferred to a 10 mL 

polypropylene beaker with clumps gently broken up, as needed, if formed during drying. This 

made about a 0.5 cm thick layer in the beaker. Epotek 301 2-FL resin (Epoxy Technologies, 

Inc) that had been outgassed in the anaerobic chamber was mixed and poured onto the dried 

sediment covering the sediment with at least another 0.5 cm layer. The beakers were then 

placed in the anaerobic chamber airlock which was then evacuated to -10” Hg to remove any 

gas phase entrained within the dried sediment. The resin cured at room temperature over the 

next three days with additionally vacuum applied to maintain desired pressure. After hardening, 

the resin pucks are removed from the beakers, labeled, and sealed in Mylar bags with oxygen 

scrubbers for shipment for thin section fabrication by Spectrum Petrographics (Vancouver, WA). 

Fabrication of thin sections entailed vertical slicing of the pucks and recast the slices into larger 

blocks using the same room temperature curing resin. Slices were cut from the block and 

mounted and polished using low oxygen and heat methods. The completed thin sections were 

transported in Mylar bags and stored in the anaerobic chamber until measured on the SSRL Xray microprobes.

Thin sections were imaged using a flatbed scanner with the scanned image enlarged to 

provide a location map. The entire area of each thin section was mapped at SSRL BL10-2 

using a 20-μm nominal beam size focused through a capillary tube using 30 μm steps and 50 

msec dwell time. Fluorescence data were collected at each pixel as the sample was stepped 

under the beam at energies of 17,100 and 17,200 eV. A difference map was then constructed 

using SMAK, the Microprobe Analysis Tool Kit (http://home.comcast.net/~sam_webb/smak.html) 

to remove the contribution of Rb fluorescence to the U fluorescence window. Single and multielement XRF maps depicting relative concentration were then constructed for each thin section 

to depict the distribution of U, Fe, Ca and other elements of interest. These maps also are used 

to locate areas for more detailed mapping at BL 2-3.

Selected areas of thin sections with higher U concentration were imaged using the SSRL 

BL2-3 X-ray microprobe using a 1 μm focused monochromatic X-ray beam with fluorescence 

data collected at 50 to 200 msec dwell time as the sample was translated across the beam. 

The effective pixel size was of 2 x 2 μm in most cases, and 5 x 5 μm for larger grains. The U 

enriched areas of the thin sections mapped with the X-ray microprobe were either individual 

grains or coatings on grain exteriors. Maps were collected at multiple energies across the U-LIII 

absorption edge at 17,170, 17,175, 17,178, and 17,190 eV. The resulting maps were 

processed using a least squares fitting routine in the SMAK software to calculate the fraction of 

U(IV) and U(VI) at each pixel based on the contribution of each component to the normalized 

fluorescence yield at these energies in XANES spectra of model U(IV) and U(VI) compounds. 
3-26

This XANES mapping routine is further described in Sharp et al (2011) and (Mayhew et al, 

2011). U-LIII XANES spectra were then collected at specific points using a 1 μm beam spot size 

on thin sections that had sufficient U to provide usable data. These points were chosen based 

on apparent distribution of U(IV) and U(VI) from the XANES maps, and locations optimized for 

maximum signal.

3.3 Results

3.3.1 Batch Experiments

The initial batch experiment with the Kingsville Dome in situ leached sediment (KVD 3712 

BC 580-590) was conducted starting in May 2010, to test the uptake of dissolved uranium under 

anaerobic conditions to determine if amendment with electron donors would result in enhanced 

uptake putatively from the reduction of U(VI) to insoluble U(IV) by the indigenous microbial 

population in the aquifer material. The results were used to determine the electron donors to 

use in biostimulation column experiments. In the absence of electron donor, no significant 

change in dissolved U(VI) concentration (20 μM) was observed over the 60 day duration of the 

experiment (Figure 3.5). In all experiments dissolved U increased to about 0.5 μM over the 24 

hour pre-equilibration period before addition of U(VI) and electron donors. In the control bottles 

(no added U or electron donor), dissolved U(VI) increased to 0.7μM over the next 4 weeks. The 

observed release of dissolved U(VI) likely was the result of desorption from surface 

complexation sites of the aquifer sediments. In the bottles with no electron donor amendment, 

no measurable change in 20 μM dissolved U(VI) ([U]) was observed over the duration of the 

experiment indicating no significant adsorption of U(VI) under the experimental conditions. 

Dissolved uranium decreased starting at about 3 days after addition in the H2 amended bottles 

and continued through 28 days, after which time a near constant [U] of 0.7 μM was measured 

that is similar to the [U] control bottle. The sediment in the hydrogen amended bottles turned 

black in color after 28 days likely because of FeS precipitation suggesting bioreduction of both 

sulfate in the AGW and iron in the aquifer sediment. Dissolved iron in bottle H1 was 7.6 μM at 

16 days compared to 2.8 μM in the control bottle (Table 3.4). Dissolved Fe had decreased to 

1.9 μM by the next sample at 28 days, the time when the dark color was first observed. 

Dissolved iron in the H2 experiment bottle was 71 μM at 16 days and had decreased to 2.4 μM 

by the next sampling. These results are consistent with bioreductive dissolution of ferric iron in 

sediments with the subsequent decrease attributed to FeS precipitation.
3-27

Figure 3.5 ISR sediment uranium batch uptake experiment 1. Dissolved uranium (μM) versus 

time for different electron donor amendments. Data points are for each batch bottle duplicate. 

Control has no added electron donor or dissolved U(VI). The “No electron donor” bottle has 20 

μM dissolved uranium, but no added electron donor.

Table 3.4 Total dissolved iron in batch experiment 1 samples filtrates (<0.2 μm) for duplicate 

bottles for each electron donor.

Time Dissolved iron (μM)

(hours) No electron donor Acetate Lactate Hydrogen

C1 C2 A1 A2 L1 L2 H1 H2

387 2.8 2.3 5.2 5.7 78 9.0 7.6 71

700 4.7 4.3 10 7.4 2.1 36 1.9 2.4

1176 11 11 13 13 5.2 108 5.6 6.9

The [U] in the 20 mM lactate amended bottle (L1) decreased starting at 12 days and 

continued throughout the course of the experiment with 3 μM [U] measured at the last time point 

(Figure 3.5). The sediment in this bottle also began to darken in color starting at 28 days. 78 μM 

[Fe] was measured in this bottle at 16 days and decreased to 2.1μM by the next time point at 29 

days (Table 3.4). Curiously, the [U] in the 10 mM lactate amended bottle (L2) was only slightly 

lower than in the control bottle throughout the experiment, but the [Fe] increased from 36 to 108 

μM between 29 and 49 days. This lag in [Fe] and no appreciable decrease in [U] in bottle L2 

suggest a lower biomass of iron and uranium reducing bacteria in the sediment L2 at the start of 

the experiment with iron reduction occurring at a later time compared to L1. The difference 

between the two lactate bottles also may be the result of a higher (20 mM) lactate amendment 

to bottle L1 than to bottle L2 (10 mM), which may also have resulted in the observed lag in 

reduction of L2 compared to L1.

0

5

10

15

20

25

0 20 40 60 80 100

ISR Batch Experiment 1

acetate

lactate

hydrogen

no donor

control

Dissolved Uranium (

μM)

Time (days)
3-28

The [U] in the acetate amended bottles did not change significantly from the initial 20 μM 

concentration through 35 days. After this time [U] in one bottle (A1) began to decrease over time 

and was at 5 μM at the last sampling point. No measurable decrease in [U] was observed in the 

other acetate bottle. Sediment in bottle A1 began to darken in color by 35 days while no color 

change was observed in the other bottle, A2. A small increase in [Fe] was measured in both 

acetate amended bottles compared to the unamended bottles. The observed differences in the 

replicate bottles may reflect heterogeneity in microbial biomass in the 3-g splits of ISR 

sediments used in the batch experiment. Heterogeneity is expected to be less of an issue in 

column experiments because of larger sediment mass used (~50 g).

The near complete uptake of [U] in both H2 amended bottles and in one of the lactate bottles 

in contrast to the unamended bottles with 20 μM U(VI) suggest that the uranium uptake was 

likely the result of biostimulated reduction. Abiotic reduction of uranium by H2 has been shown 

to be insignificant (Junier et al, 2009). This result combined with the observed release of 

significant [Fe] followed by decrease in [Fe] in the H2 and lactate amended bottles along with 

observed darkening of sediment over time suggests the presence of viable microbes capable of 

reducing iron, U, and sulfate that are present in the ISR influenced aquifer sediments. Dissolved 

sulfide was not measured in this experiment. The variability in the results among duplicate 

bottles for both lactate and acetate amended bottles suggests heterogeneity in microbial 

biomass in the 3 gram splits of sediments used in these batch experiments. Based on the 

apparent reduction of U, iron and sulfate with both hydrogen and lactate it was decided to use 

both hydrogen and lactate as electron donors in two sets of column experiments to investigate 

biostimulated reduction of U(VI) by the indigenous microbial biomass in the ISR sediment. 

Assuming sufficient biostimulated reduction and sequestration of U(IV) occurred, these columns 

also would be used to test the remobilization of uranium in response to suboxic conditions 

following cessation of electron donor. 

A subsequent set of batch experiments was conducted to test the viability of sediments in 

response to hydrogen amendment and test the effect of different dissolved H2 concentrations on 

U uptake and reduction. This experiment was initiated in June 2011 after no appreciable 

decrease in [U] was observed in the H2 amended column experiment after 3 months of flow (see 

H2 amended columns in Section 3.3.2.1). This batch experiment used the same sediment used

in both the column experiments and the first batch experiment, and at the same solid to AGW 

ratio (50 g/L). Both the 3.5% and 15% CO2 AGW were used to test for differences resulting from 

the different pH (pH 6.9 and 6.3, respectively). A range of dissolved H2 also was tested that 

spanned the H2 partial pressure imposed on the column reservoir (5%), to the level added to the 

first batch experiment (~40%), and an intermediate level (20%). These H2 partial pressures in 

the headspace resulted in equilibrium dissolved H2 of 40, 312, and 156 μM, respectively. The 

control bottles (20 μM [U] initial, no H2) had no measurable change in [U] over the 60 day 

experiment (Figure 3.6). Near complete uptake of [U] was observed for all levels of H2

amendment and for the different AGW. In bottles with 40% H2 in the headspace, a slightly 

faster decrease in [U] was observed with 3.5% CO2 AGW than with the 20% CO2 AGW. Longer 

lag times in [U] uptake were observed for a lower H2, with near complete uptake observed by 62 

days for the 20% H2 and 70% uptake for the 5% H2 amendment by 62 days. Similar to the first 

batch experiments, dissolved Fe was low in the control bottles throughout the experiment. [Fe] 

increased to 18 μM in 3.5% CO2, 40%H2 bottles through 20 days, and then decreased to near 
3-29

detection by 62 days. In contrast, [Fe] in the 15% CO2 AGW increased to between 130 and 145 

μM over the first 35 days for all three H2 amendments (Figure 3.6). [Fe] in the 20 and 40% H2

amended bottles then decreased to <1μM by day 62, with the 5% H2 bottle [Fe] decreasing to 

about 10 μM with little subsequent change by the end of the experiment. 

Total dissolved sulfur, measured by ICP in acidified samples, was equal to the AGW sulfate 

through day 36 for all bottles. The ICP sulfur is likely dominated by sulfate since samples were 

acidified several weeks prior to ICP analysis resulting in a loss of at least some of the dissolved 

sulfide as H2S gas. Dissolved S in the H2 amended bottles decreased after this time with the 

greatest decrease observed for the 3.5% CO2 AGW 40% H2 condition. Dissolved sulfide at the 

end of the experiment increased with increasing H2 in the 15% CO2 AGW with concentrations of 

0.04, 0.7, and 2.2 mM for the 5%, 20 and 40% H2 headspace amendments, respectively. The 

dissolved sulfide in the 3.5% CO2, 40%H2 bottles was 3.9 mM at end of the experiment. Batch 

bottles using sediment sterilized by gamma irradiation (25 kGy 137Cs) 10 months prior to the 

batch experiment had similar results for U, Fe and total dissolved S (data not shown) indicating 

that either the sterilization was incomplete or bacterial spores were still viable (Tuominen et al, 

1994) since abiotic sulfate reduction is not known to occur except at very slow rates.

The results of this batch experiment indicate that the microbial biomass in the archived ISR 

sediment is still viable and able to reduce U, Fe, and sulfate. Greater reduction is evident with 

the 40% H2 amendment in both AGW recipes. The sulfate reduction rate may be greater for the 

higher pH 6.9 (3.5% CO2 AGW) condition as indicated by greater decrease in total dissolved 

sulfur. The conditions of the column AGW influent (5% H2, 15% CO2) showed significant 

reduction of U and Fe but had a lag of about 30 days prior to onset of significant U reduction. 

This result is in contrast to the hydrogen columns where no measurable decrease in [U] or 

release of iron occurred over 70 days. Possible causes of this difference are discussed below. 
3-30

Figure 3.6 ISR sediment uranium batch uptake experiment 2. (A). Dissolved uranium as ratio of 

initial total dissolved (20 μM) and (B) dissolved iron versus time for different headspace H2

partial pressure amendments as electron donor, and headspace pCO2 to control pH. Data 

points are for each batch bottle duplicate. Control has no added electron donor or dissolved 

U(VI). No electron donor bottle has 20 μM dissolved uranium added, but no hydrogen added to 

headspace. IR is gamma irradiated.

 

A

B
3-31

3.3.2 Column Effluent

The following section describes the column effluent chemistry over the course of the 

biostimulation and sub-oxic elution experiments. 

3.3.2.1 Reduction stage

H2 Amended Columns. Two columns were packed with ISR sediment on February 24, 

2011. Flow of H2 amended AGW was started on March 10, 2011. The influent reservoir was 

equilibrated with a 5% H2 partial pressure which yields an equilibrium dissolved H2

concentration of 40 μM. The AGW influent was also amended with ammonium, phosphate, trace 

elements and vitamins (see Methods) to promote microbial activity and growth. An initial flow 

rate of 0.05 ml/minute was used for the first 1.2 days, a volume equivalent to 6 column pore 

volumes and residence time of 0.2 days. The flow rate was subsequently decreased to 0.012 

ml/min to increase residence time and decrease back pressure in the column. 

Effluent [U] increased to 15 μM in the second sample (0.4 to 3.7 PV) and equaled the 

influent [U] of 20 μM after 8 PV (data not shown). The lower [U] in the initial samples likely 

represents dilution with AGW without [U] used for packing the column, and adsorption of U(VI) 

to surface complexation sites on the aquifer sediment surfaces. Subsequently, no significant 

change in effluent [U] relative to influent was observed. Effluent dissolved Fe ([Fe]) and Mn 

([Mn]) increased over the first 5 days then decreased through 50 days. The release of [Fe] and 

[Mn] from columns H1 and H2 was similar in timing and extent to the effluent [Fe] and [Mn] 

observed in the lactate columns before the onset of iron reduction (data not shown). The 

decrease of [Fe] in both the hydrogen columns between 10 and 65 days suggest that no 

significant reductive dissolution of iron oxides, such as by dissimilatory Fe reduction, was 

occurring in the columns, contrasting the large increase in [Fe] observed in the lactate columns. 

Results of the H2 amended batch experiments suggested U reduction had at most a lag of 20 

days before measurable decrease in [U] was observed, despite a significantly lower solid to 

liquid ratio in the batch bottles (50 g/L) than in columns (3500 g/L), and, therefore, lower 

biomass.

Because of the lack of significant [U] uptake over the first 20 days and decrease in [Fe], 

measurement of dissolved H2 in both column effluent and in the influent at the pump outlet was 

attempted using a thermal ionization detector. Samples were collected in N2 purged 2-cc 

vacutainers with syringe needles as inlets. Displaced gas volume was vented through a syringe 

needle connected to an air-lock. The volume of the water collected was determined by weight 

and the fractional loss of dissolved H2 from headspace displacement accounted for. No H2 was 

detected in either type of sample despite a continuous 5% H2 in gas stream through both the 

reservoir and tubing jacketing, which exited near the effluent sample outlet. It was suspected 

that gas exchange across pump tubing walls resulting from pressure from pump rollers caused 

loss of dissolved H2 from the influent stream. To eliminate the loss of H2, an acrylic box was 

fabricated to house the peristaltic pump and tubing (Figure 3.4E). This box was continuously 

purged with the experimental gas mixture. This purged pump housing eliminated the problem 

with gas exchange across pump tubing and the measured H2 in both the influent and effluent 

streams became consistent with an equilibrium dissolved H2 of 40 μM. 
3-32

The H2 amended columns were then operated for an additional 60 days, with no measurable 

decrease in dissolved U or sulfate, and no significant increase in dissolved Fe observed, all 

indicating that no significant microbial reduction was occurring in the aquifer sediments. This 

time period exceeded the lag time before the onset of measurable change in dissolved U in 

batch systems with the same hydrogen partial pressure (30 days, Figure 3.6). Additional 

effluent samples for dissolved H2 indicated that H2 was entering columns but no measurable 

loss of H2 occurred within the columns. No measurable decrease in [U], increase in dissolved 

iron, or decrease in sulfate, were observed; all indicators of biostimulated microbial reduction 

expected for amendment with hydrogen electron donor. This suggests that reduction by the 

ambient microbial biomass was occurring at too low a rate to lower solution concentrations of 

these constituents in effluent at the experimental flow rates. Stopping influent flow for 24 hours 

and restarting also did not produce measurable changes in these constituents. Because of the 

apparent lack of significant microbial reduction, the H2 amended columns were stopped. These 

column experiments contradicted the observed U reduction in batch experiments. The 

sediments in batch systems were in contact with the same volume of solution for the entire 

duration allowing progressive changes in dissolved concentration to be observed whereas the 

pore volume of the column had a residence time of about 1 day so that only changes in 

concentration of the influent that occur over 1 day would result. Because no organic carbon 

source was added in the influent stream to the column there was likely no increase in microbial 

biomass during the column experiments. It is concluded that under the conditions of the column 

expeirment H2 amendment alone is not suitable for sustaining biostimulated reduction to 

significantly lower dissolved uranium in the Kingsville aquifer following ISR operations. 

Lactate Amended Columns. The flow of lactate amended AGW through columns L1 and 

L2 commenced on February 24, 2011. Significant reduction of Fe and U was evident in effluent 

profiles (Figure 3.7A). Because back pressure at the column inlet increased over the first 15 

days, the influent flow rate was decreased from 0.05 to 0.011 ml/min. Back pressure continued 

to increase in both columns over the 83 day biostimulation experiment. Effluent dissolved U 

increased over the first 2 pore volumes to the influent 20 μM concentration (Figure 3.7A). The 

slow increase likely was due in part to dilution by initial pore water from column packing that had 

no [U] and to adsorption of U(VI) by aquifer sediment. The latter may be minimal since no 

measurable uptake of U(VI) was observed in batch experiments. Effluent [U] was constant and 

about equal to the influent [U] over the first 8 pore volumes (8 days). Subsequently, effluent [U]

started decreasing and was <0.05 μM by day 50 (62 PV) indicating significant uptake, likely by 

reduction. The [U] effluent curves for columns L1 and L2 were very similar. Effluent dissolved 

iron (operationally defined as ferrous iron, [Fe]) initially was 10-15 μM and then began to 

increase starting at about 10 days and reached its’ maximum of about 300 μM in both columns 

(Figure 3.7A). [Fe] in column L1 reached its’ maximum concentration after 32 days and then 

decreased to <5 μM by day 48. [Fe] in column L2 reached maximum effluent concentration by 

day 22 then decrease to <5 μM by day 48. 

Effluent lactate decreased by 30% after 6 days and continued to decrease to near the 

detection limit by day 18, with little or no measurable lactate in the effluent through the 

remaining duration of the biostimulation stage (Figure 3.7B). This high rate of lactate 

consumption is consistent with significant microbial activity with a likely concomitant increase in 

biomass. Acetate, an oxidation product of lactate, increased from 0 (the influent concentration) 
3-33

to about 3 mM in both columns by day 12. Acetate was then relatively constant through day 55 

after which it increased and approached the influent lactate concentration (10 mM) during the 

last 20 days of the biostimulation stage. These results suggest that the acetate produced by 

lactate oxidation is in turn utilized by other microbial processes within the column. The 

consumption of acetate diminished (effluent acetate increased) at about the time effluent [Fe] 

decreased to low levels suggesting that acetate consumption may be linked to ferric iron 

reduction. Dissolved sulfate decreases at a linear rate over the entire biostimulation period 

(Figure 3.7B). Measurable dissolved sulfide in the effluent was first detected at day 6 but was 

low (<0.5mM) through day 37 and day 45 in columns L1 and L2, respectively, after which 

dissolved sulfide increased irregularly through to the end of the biostimulation experiment. The 

timing of the increase in sulfide to >1 mM was similar to [Fe] decreasing to <5 μM. This is 

consistent with iron reduction proceeding until easily reducible ferric iron is depleted followed by 

onset of significant sulfate reduction (Fang et al, 2009). 
3-34

0

5

10

15

20

25

0

50

100

150

200

250

300

350

0 40 80 120 160 200

L1 U red

L2 U red

L2_U_ox

L1 Fe red

L2 Fe red

Dissolved U(VI) 

μ

M

 Dissolved Fe (

μM)

Time (Days; 1 Pore volume = 1.06 days)

Anoxic influent

with lactate and U(VI)

Columns L1 and L2 Effluent Concentration

Lactate

Off

Suboxic Influent Dissolved Oxygen

Bromide

Tracer

"0" Oxygen

and U

Influent

6 μM 

Dissolved

Oxygen

12 μM 

Dissolved

Oxygen

0

5

10

15

20

25

0

50

100

150

200

250

300

350

0 20 40 60 80

Dissolved U(VI) 

μ

M

 Dissolved Fe (

μM)

Time (Days; 1 Pore volume = 1.06 days)

Anoxic influent

with lactate and U(VI)

Columns L1 and L2 Effluent Concentration

Figure 3.7 Columns L1 and L2 Effluent Concentrations. Part A. Column L1 and L2 effluent 

dissolved uranium and iron concentrations versus days since start of lactate addition for the 

duration of suboxic elution of column L2 (top), and expanded time scale for biostimulated 

reduction (bottom).
3-35

Figure 3.7 Columns L1 and L2 effluent Concentrations Part B. lactate and acetate 

concentrations (top) and dissolved sulfate, total dissolved sulfur (ICP) and sulfide 

(bottom) for the biostimulated period of the column experiments.

0

2

4

6

8

10

0

2

4

6

8

10

0 20 40 60 80

L2 Lactate

L1 Lactate

L2 Acetate

L1 Acetate

Dissolved Lactate (millimolar)

 Dissolved Acetate (millimolar)

Time (Days; 1 Pore volume = 1.06 days)

Columns L1 and L2 Effluent Concentration

0

2

4

6

8

10

12

0

1

2

3

4

5

0 20 40 60 80

L1 SO4

L1 Total S

L2 SO4

L1 Sulfide

L2 Sulfide

Dissolved Sulfate (millimolar)

 Dissolved Sulfide (millimolar)

Time (Days; 1 Pore volume = 1.06 days)

Columns L1 and L2 Effluent Concentration
3-36

Column L1 effluent alkalinity increased over the first 15 days of biostimulation from 6.3 

meq/L in the influent to almost 10 meq/L. Subsequently, effluent alkalinity decreased over the 

next 15 days to about 8 meq/L, after which it slowly increased to 9 meq/L by day 50 with little 

change for the remainder of the biostimulation period (Figure 3.7C). Effluent pH increased 

about 0.2 pH units and varied little during the biostimulation period. Effluent alkalinity and pH 

for column L2 were similar to column L1 except for the last two sampling points. The increase 

in alkalinity is consistent with microbial oxidation of lactate and acetate. No significant change in 

effluent dissolved Ca was observed suggesting that precipitation of CaCO3 in response to 

increased dissolved carbonate and increased pH was not sufficient to lower dissolved Ca by a 

measurable amount. Analysis of column sediments for inorganic carbonate by CO2 evolution on 

acidification showed no measurable change in total carbonate content indicating that CO2

produced during lactate and acetate oxidation was transported out of the column and not 

precipitated as CaCO3. 

Figure 3.7 Columns L1 and L2 effluent concentrations. Part C. alkalinity and pH versus 

time for the biostimulation period of the column experiment.

The effluent water chemistry data were integrated to estimate total loading of uranium, 

reduction and dissolution of iron, loss of sulfate, production of sulfide, and production of total 

dissolved carbonate. For U, sulfate, and lactate loss (or consumption) was calculated for each 

sample by subtracting the measured effluent concentration from the average measured 

concentrations in the influent reservoir. The change in concentration was then multiplied by the 

volume in liters of effluent sample collected. For total dissolved carbonate (TCO2), the 

production during each sample collection period is the difference between the measured 

concentration and the average influent concentration measured in reservoirs multiplied by the 

volume of AGW collected for each sample. The change in concentration of samples not 

0

2

4

6

8

10

5

5.4

5.8

6.2

6.6

7

0 20 40 60 80

L1 Alkalinity

L2 Alkalinity

L1 pH

L2 pH

Alkalinity (milliequilvaent/L)

pH

Time (Days; 1 Pore volume = 1.06 days)

Columns L1 and L2 Effluent Concentration
3-37

analyzed was set to the average of adjacent measured samples and then multiplied by its 

respective sample volume. The resulting change in each constituent in units of millimoles 

(micromoles for U) in each sample was then summed over the duration of the biostimulation. 

Net acetate produced was determined by integrating the mass measured in effluent samples. 

Acetate consumed during the experiment was assumed to equal the difference between lactate 

consumed and net acetate produced. The results of the total mass of production or loss of these 

constituents from the effluent integrations are summarized in Table 3.5. 

Table 3.5 Integrated loss and gain from column effluent concentrations during lactate 

biostimulation. U, sulfate, and lactate losses are the difference between influent and effluent 

concentration multiplied by sample volume summed over the biostimulation period. Acetate 

produced (net) is millimoles transported out of the column. Acetate consumed is the difference 

between lactate consumed and effluent acetate. Total amount of U, reduced Fe, and total S in 

sediments recovered from columns are the measured concentrations minus pre-column 

sediment concentration in each subsectioned intervals times mass of dry sediment recovered. 

L2 sediments were analyzed following suboxic elution. Units are all millimoles except for 

uranium, which is in micromoles.

Constituent L1 effluent L1 sediment L2 effluent L2 sediment

U uptake 17.0 16.0 16.9 13.9

U released (suboxic) 4.6 2.1

Fe reduced 0.164 0.303 0.15 0.197

Fe released (suboxic) 0.03

Lactate consumed 12.4 11.96

Acetate produced (net) 6.47 5.42

Acetate consumed 5.94 6.72

Total CO2 produced 5.1 4.6

Sulfate reduced 1.4 1.6

Sulfide produced 1.9 0.9

S retained 4.65 3.16

The calculated loading of U was divided by the total dry mass in the column and plotted as 

micrograms U per gram solid versus time (Figure 3.8). Normalizing U uptake by total dry 

weight assumes uniform distribution of U uptake by the entire sediment mass. The normalized U 

loading versus time was nearly identical in columns L1 and L2. The normalized total U uptake at 

the end of the biostimulation stage was about 70 μg U/g in each column.

Comparison of the integrated sulfate decrease with sulfide increase indicates a greater 

amount of sulfide produced than the integrated decrease in effluent sulfate. This comparison is 

based on effluent sulfide only and does not include solid phase sulfide, such as FeS 

precipitation. Inclusion of solid phase sulfide summed over the entire column sediment mass 

results in a factor of 4 or more S reduced than can be accounted for in the decrease in 

dissolved sulfate (see below). One possibility to explain this difference is that oxidation of 

sulfide occurred in the effluent sample bottles between collection of the effluent samples used 

for sulfate analysis by ion chromatography analysis. These samples were collected in sterile 
3-38

serum bottles that had been flushed with the zero oxygen gas mixture into which effluent flowed 

through a 0.22 μm filter with headspace vented through an airlock. Little difference is observed 

between total S measured by ICP in acidified effluent samples and sulfate in unacidified effluent 

samples measured by ion chromatography when both are plotted versus time or volume 

passed. This comparison suggests that the sulfide in the unacidified samples had oxidized 

either during storage or in preparation for sulfate analysis. The decrease in total S and sulfate 

over time in the effluent is an indicator only of S removed from the influent by precipitation of 

sulfide in the column and thus not a measure of total sulfate reduction.

Figure 3.8 Cumulative loading of uranium on column L1 and L2 sediment versus time of the 

experiment. U loading during biostimulation (Ured) is calculated by the difference between the 

influent and effluent uranium times sample volume and divided by total dry mass of sediment. 

Loading is summed up to a given sampling time

 Unmeasured sampling intervals are set equal to average of adjacent intervals. Uranium 

loading during the suboxic elution (Uox loading) for column L2, is the total loading at end of 

biostimulation minus the sum of the product of effluent uranium times the sample volume 

divided by the total dry mass of sediment.
3-39

3.3.2.2 Suboxic elution - extent of uranium remobilization

The near complete removal of [U] from the influent over the course of the biostimulation 

stage in the lactate columns suggests that the stimulation of ambient microbial population in the 

Kingsville Dome aquifer after in situ recovery of U may be an effective means of remediating 

dissolved uranium in groundwater following the extraction process. The removal of dissolved 

uranium, which is in the +6 oxidation state, likely occurred predominantly by reduction to U(IV) 

and precipitation, since the +4 oxidation state has much lower solubility than U(VI) for the 

groundwater chemistry of the Kingsville Dome site. Measurement of the U oxidation state and 

characterization of the form of the U removed during biostimulation is presented in section 3.3.4. 

Determining the stability of the bioreduced uranium in response to changes in groundwater 

chemistry, such as presence of dissolved oxygen at suboxic levels, is needed to evaluate if the 

biostimulation process would be an effective tool for long term remediation of ISR influenced

aquifers. Pre-operational dissolved oxygen concentrations are not well known with most 

reported values below the working range of dissolved oxygen meters (0.5 mg/L). A lower value 

of 0.2 mg/L (6 μM) was chosen to test suboxic conditions typically found in deep aquifers. 

The stability of bioreduced U was tested by elution of column L2 after biostimulation by 

flowing AGW with no [U] or lactate through the column (see above). Prior to suboxic elution but 

after biostimulation, a bromide pulse was passed through column L2. A reservoir containing 

bromide (360 mg/L) in AGW (no U, lactate, vitamins or nutrients) equilibrated with 15% CO2/N2

gas mixture was flowed through the column for 5 pore volumes (days 85 to 90) to determine the 

pore volume at the end of biostimulation. Following the bromide pulse, a reservoir with AGW 

with no bromide, U, or lactate was equilibrated with 15% CO2/N2mixture (no O2) and flowed 

through the column to elute bromide. After seven 4-hour samples, the effluent collection time 

was increased to 6 hours with effluent samples collected over the next 4 days to capture the tail 

of eluted bromide. Elution of bromide was sampled for a total of 7 days. Splits of every fourth or 

fifth sample during the bromide pulse and tail sampling were acidified for [U] analysis. Elution of 

the column with this U-free AGW equilibrated with 15% CO2/N2 gas mixture (with no added 

oxygen) continued for the next 32 days (June 3 –July 8, 2011, day 101 to 133), with samples 

collected in serum bottles at 2 to 3 day intervals. Subsequently (starting July 8, day 133), 

oxygen at 0.5% was included in the gas mixture entering the reservoir and jacketing to attain an 

equilibrium dissolved oxygen concentration (DO) of 0.21 mg/L (6 μM). A high flow rate of the 

gas mixture was used for one hour to equilibrate the reservoir with this new gas mixture. On day 

151, three weeks after the start of suboxic elution, measurement of dissolved oxygen in the 

column effluent (see method above) yielded 1.4 mg/L (44 μM). The outflow sampled directly 

from the reservoir upstream of the peristaltic pump yielded a DO of 0.2 mg/L. These 

measurements indicated that atmospheric oxygen was entering the influent in the pump tubing 

as a result of the pumping process. Dissolved oxygen entering the influent during biostimulation 

likely was consumed rapidily by aeorphillic bacteria near the column inlet and did not affect 

influent latacte concentration. On day 153, the pump was installed within the acrylic housing 

that was purged with the same 0.5% O2/15%CO2/N2 mixture. The resulting DO in both pump 

outlet and in the column effluent was 0.4 mg/L (12 μM), which remained constant throughout the 

remainder of the suboxic elution that continued for 67 more days to day 220.
3-40

During the bromide pulse and elution, and prior to addition of oxygen to the influent 

reservoir, effluent U was less than 0.01 μM (<3 μg/L) (Figure 3.9). Dissolved Fe fluctuated 

irregularly between 1.4 and 7.8 μM during the bromide pulse and tail. Starting at day 101, 

effluent Fe increased rapidly to 18 μM, and subsequently varied between 18 and 24 μM over the 

remaining 30 days prior to introduction of O2 in the gas stream. Following introduction of O2 to 

the reservoir and jacketing gas at 0.5% at day 133, the effluent U increased rapidly to 0.1 μM 

(20 μg/L), within a couple pore volumes, and then remained near this level over the next 5 days. 

From this point onwards (day 133), effluent samples were collected using the fraction collector 

to facilitate shorter collection periods. After day 140, the effluent U concentration started 

increasing, with the rate increasing with time, reaching 0.13 μM (30 μg/L) by day 144 and 0.25 

μM (60 μg/L) by day 150. Effluent dissolved Fe decreased somewhat linearly from day 120 

onwards, but decreased to a minimum between day 133 and day 138. This decrease in Fe that 

coincided with use of the fraction collector was likely due to oxidation and precipitation in the 

tubing connecting the column effluent to the fraction collector. The effluent line to the fraction 

collector was jacketed and purged with the gas mixture starting at day 138, which resulted in 

effluent Fe increasing to about the concentration measured prior to use of the fraction collector, 

about 20 μM. Subsequently, effluent [Fe] decreased throughout the remainder of the experiment 

and was about 6 to 7 μM by the end of the experiment at day 220. The cumulative Fe 

transported out of the column in the effluent during suboxic elution is attributed to oxidative 

dissolution of reduced iron phases, like iron sulfides, with some of the ferrous iron transported 

out the column prior to significant oxidation to ferric iron because of the well-known effect of

carbonate and sulfate complexation of ferrous iron slowing the rate of ferrous oxidation in AGW.

Release of U, likely from oxidation and dissolution, increased throughout the suboxic elution 

with the overall rate of release increasing greatly after about day 180. This is evident from the 

increase in effluent U from 0.5 μM at day 180 to over 12 μM by day 220, when the experiment 

was terminated. Effluent Fe decreased from about day 150 through the end of the experiment. 

The decrease in effluent Fe was approximately coincident with the increase in the rate of 

effluent [U] suggesting that the apparent increase in the rate of U oxidation may be the result of 

depletion of iron sulfides which are competing for dissolved oxygen. Very little change in 

effluent sulfate was measured during the suboxic elution indicating that sulfides did not oxidize 

completely to sulfate and/or that sulfate from sulfide oxidation was not readily released to 

solution.
3-41

Figure 3.9 Column L2 effluent dissolved uranium (red triangles) and iron (blue diamonds) 

during suboxic elution versus days since start of lactate addition. This is an expanded scale for 

Fe and time from Figure 3.7A. Effluent samples were collected with a fraction collection 

throughout the suboxic elution except for the period between 97 and 132 days, when samples 

were collected into serum bottles with outlet lines fully jacketed with equilibration gas. Oxygen 

(0.5%) was added to equilibration gas at day 133. Outlet line to fraction collector was jacketed 

with equilibration gas starting at day 138. Low Fe in effluent between days 133 and 138 was 

likely because of oxidation and precipitation in outlet line. The pump housing with equilibration 

gas flush was installed at day 153.

The integrated mass of U remobilized was determined on a per gram basis over the duration 

of the oxidation stage for column L2 from the effluent [U]. The mass of U in each effluent sample 

(concentration times sample volume) was divided by the total dry sediment weight in the column 

and then subtracted from the calculated total U loading during the biostimulated reduction 

period (Table 5 and Figure 3.8). By day 180 about 2% of the total cumulative U uptake during 

biostimulated reduction had been remobilized, with more than 26% remobilized by the end of 

the experiment. The U loading in excess of the initial sediment U content over the course of the 

reduction and sub-oxic elution stages is shown in Figure 3.8. Although the experiment was 

ended after 220 days, the trajectory of both the effluent dissolved U and U loading after day 180 

suggests that most if not all of U uptake during reduction would be remobilized from the 

sediment into solution, even at dissolved oxygen concentrations of 0.4 mg/L. The entire mass 

of U uptake is estimated to be eluted by 160 days (150 pore volumes) after the start of suboxic 

elution (experiment day 133) by extrapolation of U mass versus time in Figure 3.8. 

A fifth column of Kingsville sediment, L5, was not subject to bioreduction. This column was 

used to determine the potential for re-mobilization of the remaining uranium associated with the 

aquifer sediment (57±7 μg/g or 0.24±0.03 μmol/g) following in situ leaching. The results of this 
3-42

column are used to estimate the contribution of this source of U to the observed mass of U 

remobilized during suboxic elution of column L2. The effluent profile for the elution of column L5

is shown for dissolved U and Fe (Figure 3.10). The profile of U eluted from column L1 during 

the biostimulation stage is shown for comparison. The column pump was contained within the 

acrylic housing with continuous gas purging throughout the duration of the experiment, initially 

with no oxygen (15% CO2/balance N2), and subsequently with oxygen included (0.5% 

O2/15%CO2/balance N2) for a 0.4 mg/L (12 μM) dissolved oxygen starting at 11 PV (day 13). 

After an initial 1.6 PV of flow with AGW with no dissolved U or added oxygen, AGW with 

dissolved bromide was flowed from PV 1.6 to 6.5 (7.8 days). Subsequently, Br tracer was 

eluted with Br-free AGW, also with no added dissolved oxygen. Effluent Br was monitored for 

an additional 6 PV to day 15.

Effluent [U] was 5μM over the first 3 days, then decreased to <2 μM over the next 2 days. It

continued to decrease through the duration of the elution, although a small increase to 1.1 μM 

occurred between 7 and 12 PV, which coincided with change to bromide-free AGW, but

preceded introduction of dissolved oxygen at 12 PV. Dissolved uranium decreased through 43 

PV at which point [U] was ~0.2 μM and the experiment was ended. The total integrated U 

released from the ISR sediments to solution during the column L5 experiment was 3.0 μg/g 

(0.013 μmol/g), which is about 4.6% of the total U in the ISR sediment, as recovered from the 

field site. About 1.7 μg/g or 60% of the total U released from column L5 occurred prior to 

introduction of dissolved oxygen at 13 PV. By comparison, a much lower mass of U (0.03 μg/g) 

was released from column L2 over 44 PV during the Br tracer prior to introduction of dissolved 

oxygen. The total U released during the column L5 experiment is about 3.5% of the total U 

remobilized from column L2 during suboxic elution. These results suggest that either this 

fraction of residual U in the Kingsville sediments after ISR was either mobilized during the 

biostimulation stage of column L1 and L2, or that the biostimulation limited the release of the 

residual U during suboxic elution. 

Dissolved Fe in column L5 increased over the first 2 days to about 35 μM and was relatively 

constant until day 12. After this point and from the start of 12 μM dissolved oxygen, effluent Fe 

began to decrease exponentially to less than 2 μM by the end of the experiment, except for a 

small increase between 18 and 20 days. Since dissolved Fe in effluent (<0.2 μm filtrate) is likely 

ferrous iron, desorption of Fe2+ or dissolution of a ferrous iron phase is the likely source of 

effluent iron. A similar initial increase in effluent iron in columns L1 and L2 was observed during 

the first few days prior to large increase in effluent Fe from day 10 through 40 (Figure 3.7A). 

Column L5 was subsectioned for solid phase analyses as described for columns L1 and L2.
3-43

Figure 3.10 Column L5 suboxic AGW elution of the ISR 3712 B-C 580-590 sediment. Effluent 

dissolved uranium (blue diamonds) and iron (red triangles) versus days. Column L1 dissolved U 

(green diamonds) during biostimulated reduction with 20 μM U(VI) influent is shown for 

comparison.

3.3.3 Solid Phase Characterization of Column Sediments

The following section describes results of geochemical and microbial characterization of 

sediments recovered at the end of biostimulation (column L1), suboxic elution (column L2 and 

L5), and the sediment recovered from coring of Kingsville Dome that was used for the column 

experiments (3712 BC 580-590). 

The Kingsville Dome ISR sediment used for this study is primarily fine sand and silt (Table 

3.6) and has a N2-BET surface area of 11.4 m2

/g. Quantitative X-ray diffraction analysis shows 

that the sediment is comprised primarily of quartz, 42%, calcite, 22%, and feldspars, 21% 

(Table 3. 7), indicative of the calcareous sandstone. Clays, primarily muscovite and haloysite, 

comprise 10% of the mass. Iron, as oxides, account for 3%, and as sulfides 1.6%, as marcasite 

and pyrite. Gypsum is about 0.4% by weight. Total carbonate, as determined by acid 

volatilization, is 12.4% weight, is equivalent to 20.8±0.2 weight % as CaCO3, and comparable to 

the 22% calcite from XRD. The total S content of 0.75% measured by EA is equivalent to 2.7% 

FeS2 after accounting for S in gypsum.
3-44

Table 3.6 Kingsville Dome ISR 3712 B-C 580-590 sediment grain size distribution as 

determined by Coulter grain size analyzer.

Grain Size Weight %

microns

<20.7 15.1

20.7 - 63 17.8

63 - 122 29.1

122 - 257 31.7

256 - 494 4.7

494 - 1041 1.6

Table 3.7 Mineral weight percent of Kingsville Dome ISL sediment 3712 B-C 580-590 (precolumn) as determined by quantitative x-ray diffraction.

Mineral Weight %

Non-Clays

Quartz 41.7

Kspar (ordered Microcline) 1.5

Kspar (orthoclase) 10.2

Plagioclase (albite, var. cleavelandite) 3.0

Plagioclase (oligoclase; NC) 0.8

Plagioclase (andesine) 3.7

Plagioclase (anorthite) 1.8

Calcite 22.3

Pyrite 0.4

Marcasite 1.2

Gypsum 0.4

Ferrihydrite (Humbug Creek) 3.0

Total non-clays 90.1

Clays

Halloysite 4.9

Muscovite 5.1

Total clays 9.9

Total 100.0

3.3.3.1 Uranium

A total uranium concentration of the Kingsville ISR sediment of 57±7 μg/g (0.24±0.03 

μmol/g) was measured in the pre-column sediment (as recovered from the field site) by hot nitric 

acid extraction. This concentration indicates that >95% of the uranium had been removed by 

the in situ recovery process assuming that the sampled sediment originally contained ore of 

0.088% grade. The remaining U likely was inaccessible to ISR lixiviant and/ or not readily 
3-45

oxidized. The total U distribution in the pre-column sediment is slightly higher in finer grain size 

fractions with 74% in the <125 μm, 22% in the 125-250 μm fraction, and 5% in the >250 μm 

fraction. The fractions comprise 62, 32 and 6% of the total sediment mass, respectively.

At the end of the bioreduction stage, the total U concentration in column L1 sediments 

ranged from 230 μg/g (1 μmol/g) at the inlet to near background in the interval at the outlet end 

(70 μg/g or 0.3 μmol/g, Table 3.8A) indicating preferential U attenuation in the upstream end of 

the column. The net uptake of U for each interval (total minus the initial or pre-column U) shows 

that greater U uptake occurred near the inlet (Figure 3.11). The net increase in total U mass 

(concentration for each interval times its total dry weight of sediment recovered) for each 

interval shows a similar distribution of uptake. The fractional uptake by each interval shows that 

95% of U removal occurred over the first 6 cm of the column. The total U uptake for column L1 

(sum of the interval concentrations times mass), 16.0 μmoles, is similar to the total uptake, 17.0 

μmoles, estimated by integrating the change in effluent [U] over the biostimulation reduction 

period (Table 3.5).

Total U concentrations in sediment from column L2 ranged from 76 to 169 μg/g (0.32 to 0.71 

μmol/g), and after correcting the background U concentration, 19 to 112 μg/g (0.08 to 0.47 

μmol/g) (Table 8A). Comparison of the U distribution after suboxic elution (column L2) to before 

elution (column L1) shows that U remobilization occurred from the upstream half of the column, 

in particular, from the 2-4 cm interval, and that some U was higher near the outlet end. These 

comparisons assume that U distribution in column L2 sediment at the end of the biostimulation 

period was the same as measured in column L1. It is unknown if the higher U near the outlet is 

the result of greater uptake during biostimulation or if U was retained near the outlet end during 

suboxic elution. The measured total U in column L2 (sum of the interval concentrations times 

mass) is 13.9 μmoles. Assuming that the total U uptake during biostimulation was the same in 

both columns, the difference between solid U represents a remobilization of 2.1 μmol U, or 

about 15% of the total uptake. The U uptake in both columns calculated from the integrated 

change in effluent [U] during biostimulation was the same (Table 3.5). By comparison, 4.6 μmol

of U remobilized during the suboxic elution was calculated from the integrated effluent [U], 

which is about 27% of the 17.0 μmoles U uptake during biostimulation remobilized during 

suboxic elution. 

The sediment U content in column L5 decreased from the initial background of 13.2 μmoles 

to 12.7 μmoles during the elution with no dissolved oxygen AGW followed by suboxic AGW. 

Sediment U concentration varied little among the subsampled intervals (Table 3.8). A total U 

release of 0.56 μmol was estimated from the sum of the measured U in each sediment interval 

multiplied by its dry mass. The integrated effluent [U] over the duration of the elution was 0.74 

μmol. These estimates of U release represent 4 to 5 % of the total initial U in column L5. The 

effluent history indicates most of the release occurred during the first 8 pore volumes prior to 

introduction of dissolved oxygen. It is unknown if an equivalent amount of background U was 

remobilized during the first 8 pore volumes of the biostimulation experiment. Regardless, the 

amount of U released from the pre-column ISR sediment does not account for the differences in 

uptake and release calculated from the integrated changes in effluent [U] and the U extracted 

from the recovered sediment.
3-46

Table 3.8A Column sediment solid phase uranium concentrations and whole column mass for 

column L1 after biostimulated reduction, column L2 after suboxic elution following reduction, 

and column L5 after suboxic elution.

Column 

interval

Distance 

from 

inlet 

(cm)

Interval 

total dry 

weight 

(g)

Interval 

weight 

fraction 

of 

whole 

column

Total U 

(HNO3) 

(μg/g) 

stdeva Total U 

(μmol/g)

Change 

from 

pre-col 

U 

(μmol/g)

Total U 

interval 

(μmol) 

Change 

in U 

interval 

(μmol) 

Fraction 

of 

uptake 

by 

interval

L1-R1 0-1 6.22 0.112 230 0.97 0.73 6.02 4.52 0.28

L1-R2 1-3 10.92 0.197 192 1.0 0.81 0.57 8.80 6.17 0.39

L1-R3 3-5 10.87 0.196 134 0.56 0.32 6.10 3.49 0.22

L1-R4 5-7.5 11.24 0.203 78 0.33 0.09 3.66 0.96 0.06

L1-R5 7.5-10 16.24 0.293 70 0.29 0.05 4.76 0.86 0.05

L1 total 55.49 29.3 16.0

Effluent total 17.0

L2-Ox1 0-2 12.66 0.232 170 0.71 0.47 9.03 5.99 0.43

L2-Ox2 2-4 9.09 0.167 77 0.32 0.08 2.94 0.75 0.05

L2-Ox3 4-6.5 13.75 0.252 93 1.7 0.39 0.15 5.39 2.09 0.15

L2-Ox4 6.5-8 8.27 0.152 121 0.51 0.27 4.20 2.21 0.16

L2-Ox5 8-10 10.75 0.197 121 0.51 0.27 5.45 2.87 0.21

L2 total 54.52 27.0 13.9

L5-1 0-2 11.07 0.201 55.3 0.2 0.23 0.01 2.57 0.09

L5-2 2-4 9.11 0.165 53.6 0.23 0.01 2.05 0.14

L5-3 4-6 11.51 0.209 53.6 0.22 0.02 2.59 0.17

L5-4 6-8 11.07 0.201 55.9 0.23 0.01 2.60 0.06

L5-5 8-10 12.32 0.224 55.1 0.23 0.01 2.85 0.11

L5 average 0.23 0.01

L5 total 55.09 12.7 0.57

Effluent total 0.74

ISR pre-column sediment 57.2 7.0 0.24

a. Standard deviation of replicate analyses propagated through calculation.
Table 3.8B. Column sediment solid phase extractable and total iron concentrations for column L1 afte

reduction, column L2 after suboxic elution following reduction, and column L5 after subox

Column 

interval

distance 

from 

inlet 

(cm)

1 hr 

HCla

Fe(II) 

(μmol/g) stdevd

increase 

Fe(II) 

from 

pre-col

1 hr HCl 

Total Fe 

(μmol/g) stdevd

1 HR 

HH Feb 

(μmol/g) stdevd

Reduciblec 

HH FeFe(II)

Fe(II) 

increase 

interval 

(μmol)

L1-R1 0-1 10.3 7.0 10.2 11.8 1.0 1.4 43.8 

L1-R2 1-3 8.9 0.1 5.7 8.7 0.1 9.6 0.9 0.7 61.7 

L1-R3 3-5 8.7 5.4 8.6 9.8 0.8 1.1 58.5 

L1-R4 5-7.5 8.1 4.8 8.1 9.4 0.4 1.3 53.9 

L1-R5 7.5-10 8.5 0.1 5.2 8.6 0.1 9.7 1.6 1.1 85.2 

L1 average 8.9 5.6 8.8 10.0 1.1

L1 total 303 

Effluent total 164

L2-Ox1 0-2 6.8 0.9 3.5 6.9 1.1 7.0 0.2 0.2 44.5 

L2-Ox2 2-4 7.4 4.1 7.3 8.0 0.0 0.6 37.4 

L2-Ox3 4-6.5 6.5 0.2 3.2 6.4 0.2 6.6 0.5 0.2 43.9 

L2-Ox4 6.5-8 7.0 3.7 7.0 8.2 0.2 1.1 31.0 

L2-Ox5 8-10 7.0 3.7 6.9 8.5 0.1 1.5 39.8 

L2 average 6.9 3.7 6.9 7.7 0.7

L2 total 196 

Difference L1-L2 107

pre-column 3.3 0.0 4.4 0.1 1.1 

a. 1 hr HCl: sediment extracted with 0.5N HCl for 1 hour at room temperature

b. HH: sediment extracted with 0.25 N hydroxylamine hydrochloride in 0.5 N HCl at room temperature for 1 hour.

c. Reducible Fe is difference between HH extractable Fe and Fe(II) measured by 0.5N HCl.

d. Standard deviation of replicate analyses propagated through calculation.

3-47
Table 3.8C. Column sediment solid phase total and acid volatile sulfur concentrations for column L1 a

reduction, column L2 after suboxic elution following reduction, and column L5 after subox

Column interval total S 

(mmoles/g) 

stdeva S 

increase 

from 

precolumn

(mmol/g)

stdeva total S 

retained 

(mmoles)b

stdeva

δ34S

total S

‰

stdeva calculated 

δ34S of S 

increasec 

‰

L1-R1 0.43 0.03 0.20 0.02 1.23 0.15 -39.31 0.38 -41.9 

L1-R2 0.33 0.02 0.09 0.02 0.98 0.26 -38.76 0.22 -43.0 

L1-R3 0.31 0.08 0.02 0.84 0.26 -38.49 -42.6 

L1-R4 0.30 0.06 0.02 0.69 0.27 -37.94 -41.0 

L1-R5 0.31 0.07 0.02 1.19 0.39 -38.86 -44.4 

L1 average 0.10

L1 total 4.93 0.61

L2-Ox1 0.30 0.07 0.02 0.87 0.30 -37.76 -38.6 

L2-Ox2 0.27 0.04 0.02 0.34 0.22 -37.68 -38.8 

L2-Ox3 0.30 0.01 0.06 0.02 0.87 0.33 -38.37 0.26 -41.6 

L2-Ox4 0.29 0.06 0.02 0.51 0.20 -38.28 -41.2 

L2-Ox5 0.30 0.07 0.02 0.70 0.26 -37.88 -39.2 

L2 average 0.06

L2 total 3.28 0.59

L5-1 0.19 -36.37

L5-2 0.32 -39.65

L5-3 0.27 -37.86

L5-4 0.29 -38.93

L5-5 0.23 0.02 -37.67

L5 average 0.26 0.05 -38.10

pre-column 0.23 0.01 -37.20 0.68 

 a. Standard deviation of replicate analyses propagated through calculation.

 b. Total S increase for interval

 c. Calculated δ34S of total S precipitated during biostimulation period

 d. AVS is acid volatile sulfur concentration of sediment. 

3-48
3-49

 

0.0

0.2

0.4

0.6

0.8

1.0

1.2

0-2 2-4 4-6 6-8 8-10

Uranium (

μmol/g)

Distance from Inlet (cm)

L1 after reduction

L2 after suboxic elution

L5 suboxic elution only

Pre-column

0.0

0.1

0.2

0.3

0.4

0.5

0-2 2-4 4-6 6-8 8-10

Fraction of Total Uranium Uptake 

Distance from Inlet (cm)

L1 after reduction

L2 after suboxic elution

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

0-2 2-4 4-6 6-8 8-10

Uranium Uptake (

μmol/interval)

Distance from Inlet (cm)

L1 after reduction

L2 after suboxic elution

A

B

C

Figure 3.11 Solid phase uranium concentrations of sediments recovered after biostimulated 

reduction period (column L1) and after suboxic elution (column L2 and L5). (A) Uranium 

concentrations in columns L1, L2, and L5 by interval with pre-column sediment shown for 

comparison. (B) Mass distribution of uranium uptake by column interval in μmoles for 

columns L1 and L2 as the increase above initial total U of pre-column sediment. (C) Fraction 

of total uranium uptake in column by interval for columns L1 and L2.
3-50

3.3.3.2 Extractable ferrous and ferric iron

Extractable ferrous iron (Fe(II)) measured in 1 hour 0.5 N HCl extraction of column L1 

sediments ranged from 8.1 to 10.5 μmol/g compared to 3.3 μmol/g Fe(II) in the pre-column 

sediment (Figure 3.12A; Table 3.8B). Fe(II) concentration was slightly higher at the inlet end. 

The increase in Fe(II) measured in the L1 column sediments indicates that either reduction of 

some of the Fe(III) occurs within the solid phase without mobilization of the Fe(II), or that 

mobilized Fe(II) precipitates within the column as a sulfide such as FeS or as siderite( FeCO3). 

The increase in ferrous iron for the whole column was calculated by subtracting the pre-column 

Fe(II) from the post biostimulation Fe(II), multiplying by the total weight of each interval and 

summing. This results in a total Fe(II) increase of 303 μmol for column L1, which is about two 

times the integrated iron release measured in the effluent (164 μmol). The increase in sediment 

Fe(II) summed with effluent Fe(II) yields a total Fe reduction of 467 μmol during the 

biostimulation, or 8.4 μmol/g. The reducible iron fraction defined by the difference in 1 hour HH 

(ferrous and ferric iron) and 1 hour 0.5N HCl (ferrous only) extractions in the column L1 after 

biostimulation ranged from 0.7 to 1.4 μmol/g. This range encompasses the pre-column sediment 

reducible Fe, 1.1 umol/g. The much greater observed Fe reduction (8.4 μmol/g) than the 

measured reducible Fe in the pre-column sediments (1.1 μmol/g) indicates that this measure of 

reducible iron is not representative of the ferric iron reduced during the biostimulation stage of 

the column experiment. A less labile or structural form of ferric iron that is reduced in place 

and/or solubilized as Fe(II) likely accounts for the difference in iron reduction. Reduction of 

Fe(III) in phyllosilicate minerals by dissimilatory microbial iron reduction can be significant in 

aquifer sediments (Wu et al, 2012; Lee et al, 2012; Komlos et al, 2007). The effectiveness of 

0.5N HCl for extraction of Fe(II) produced by reduction of Fe(III) silicates is unknown. 

After suboxic elution, sediment Fe(II) concentrations in column L2 ranged from 6.4 to 7.3 

μmol/g, about 20% lower on average than column L1 (Figure 3.12B; Table 3.8B), and varied 

little along the flow axis of the column. Fe(II) comprised most of the readily extractable iron from 

Column L2 sediments. The reducible iron concentration in column L2 had a similar range (0.2 to 

1.5 μmol/g) as column L1, but on average was about 40% lower than the pre-column sediment. 

The HH minus HCl extractable Fe also represents labile ferric Fe, which is slightly higher in the 

effluent end of column L2 (Figure 3.12). The increase of Fe(II) over the initial or pre-column 

Fe(II) sediment concentration summed over the entire mass of sediment in column L2 was 197 

μmol, compared to 303 for column L1, indicating a decrease of 107 μmol, or about 30% of the 

Fe(II), assuming a similar amount of reduction of iron in both columns. This 30% decrease in 

Fe(II) in column L2 is consistent with mobilization of iron from sediment and transport out of the 

column during the suboxic elution since there is no measurable increase in ferric iron in column 

L2 relative to L1. Low concentrations of dissolved Fe were measured in effluent during suboxic 

elution, with generally decreasing concentrations over time (Figure 3.9). Although, iron 

speciation was not measured, the effluent iron likely was predominantly Fe(II). Integrated Fe 

release from effluent (28.9 μmol), is much lower than the difference between total Fe(II) 

produced in columns L1 and L2 (107 μmol), suggesting that most of the iron (~75%) is not 

remobilized during oxidation. Instead, it is likely transformed to a form not soluble in 0.5N HCl or 

in HH. In column studies of biostimulated reduction and re-oxidation of iron in aquifer sediments, 
3-51

Komlos et al (2007) found that the mineralogy of Fe silicates was largely the same after reoxidation as before biostimulated reduction indicating that biogenic reduced Fe(II) is re-oxidized 

back to silicate Fe(III). These results are consistent with a component of re-oxidized Fe(III) that 

remains in silicate mineral structures and therefore does not contribute to extractable Fe 

reported here. 

Sediment iron concentrations measured in the hot HNO3 extraction and in HF total 

dissolution are shown in Table 3.8B. Higher iron is measured in HF dissolution than in hot 

HNO3 digests consistent with a fraction of iron in silicate minerals that HNO3 does not dissolve. 

The difference between HF and HNO3 iron is smaller after biostimulation, consistent with 

reduction of Fe(III) in silicate to HNO3 soluble form. Iron oxidation during suboxic elution 

resulted in no measurable change in this difference compared to after biostimulation.

0.0

2.0

4.0

6.0

8.0

10.0

12.0

14.0

0-2 2-4 4-6 6-8 8-10

Extractable Fe (

μmol/g)

Distance from Inlet (cm)

L1 HCl

L1 HH

Pre-Column HCl

Pre-column HH

A

0.0

2.0

4.0

6.0

8.0

10.0

12.0

14.0

0-2 2-4 4-6 6-8 8-10

Extractable Fe (

μmol/g)

Distance from Inlet (cm)

L2 HCl

L2 HH

Pre-Column HCl

Pre-column HH

B

Figure 3.12 Extractable iron concentration in column sediments by interval for (A) column 

L1after biostimulated reduction and (B) column L2 after suboxic elution, with pre-column 

sediment shown for comparison. HCl represents ferrous iron as defined by 1 hour 0.5 N HCl 

extraction in absence of oxygen, and HH represents both ferrous and ferric as defined by 1 

hour room temperature hydroxylamine hydrochloride extraction.
3-52

3.3.3.3 Total and acid volatile sulfur

The total sulfur content of pre-column sediments was 0.75% or 0.24 mmol/g. Total S in 

column L1 after biostimulation ranged from 0.3 to 0.51 mmol/g, with highest S at the inlet end of 

the column (Figure 3.13A; Table 3.8C) Correcting for the pre-column S, assuming no loss 

during the experiment, a net increase of 0.06 to 0.18 mmol/g is calculated. A total S increase in 

column L1 of 4.65 mmol is calculated by summing the increase in S in each interval times its dry 

mass. This increase in S retained in column L1 is about 2.5 times greater than integrated 

effluent sulfide mass of 1.9 mmol. Combined these equal 6.5 mmol of S and represent the total 

sulfate reduced during biostimulation in column L1. The total S increase in column sediment is 

15 times greater than the increase in sediment Fe(II) for the column (0.3 mmol Fe), and implies 

that most of the S retained is not precipitated as iron sulfides, such as FeS.

Stable sulfur isotopes measured during the total S analyses (Figure 3.13B) ranged from -

37.91 to -39.74 per mil δ34S which represents a depletion in the heavier isotope, 34S, of -0.34 to 

-2.14 per mil (‰) δ34S relative to the pre-column sediment δ34S of -37.2 ‰. The most 34S 

depleted sample was at the inflow end of the column which had the greatest increase in total S. 

Microbial sulfate reduction typically results in isotope fractionation on the order -20 or more ‰ 

δ34S (Habicht and Canfield, 1997). S isotope fractionation on this order is consistent with a 

microbial reduction process, which is expected in the biostimulation columns. The δ34S of the S 

retained or precipitated in the column was calculated based on two component mass balance,

δ34Stot * Stot = δ34Sppt * Sppt + δ34Spre*Spre, 

where, δ34Stot, δ34Sppt, δ34Spre are the sulfur isotope ratios for the total S after biostimulation, 

the S precipitated in the column, and the S initially in the sediment prior to biostimulation, 

respectively, and Stot, Sppt, Spre are the concentration of S in the sediment after biostimulation, S 

precipitated during biostimulation, and the initial total S in the sediment.

The δ34S of the precipitated S calculated from this equation ranges from -39.2 to -42.9 per 

mil, which indicates fractionion of about -40 ‰ relative to the AGW sulfate δ34S of -1.0 ‰ (Table 

3.8). By comparison, the δ34S of the total S of the column sediment would range from -21.8 to -

30.0 ‰ if no fractionation of the AGW sulfate S occurred during reduction. This was calculated 

using the above equation substituting the AGW sulfate δ34S (-1.0 per mil) for δ34S(ppt) and 

solving for δ34S(tot). The observed S isotope fractionation of -38.2 to -41.9 ‰ is consistent with 

a microbial sulfate reduction process. 

Total S in column L2 sediments ranged from 0.27 to 0.30 mmol/g, or an increase of 0.03 to 

0.07 mmol/g, with no apparent trend along the flow path (Figure 3.13). The S increase 

summed for the whole of column L2 was 3.28 mmol. Assuming the same retention or 

precipitation of S occurred during biostimulation in column L2 as measured in column L1, the 

difference of 1.65 mmol is the loss of S during the suboxic elution, likely resulting from oxidation 

of reduced S species to soluble form. However, there was no measurable change in effluent 

sulfate during the course of elution of column L2 starting with bromide tracer through to the end 

of suboxic elution. During this period a total volume of 2.03 L of AGW was flowed through the 

column. If the apparent decrease in sediment S was from continuous transport of sulfate out of 

the column, this decrease in S divided by the total volume of flow an average would result in an 
3-53

increase of 0.7 mM SO4 in the effluent. Instead, average sulfate measured in the effluent was 

10.75 mM which is much closer to the influent concentration of 10.65 mM. 

The δ34S of the column L2 sediments ranged for -37.68 to -38.37 ‰ δ34S (Figure 3.13B), for 

a depletion of -0.08 to -0.77 ‰ δ34S. The calculated δ34S of the precipitated S in column L2 

ranged from -38.2 to -41.3 ‰. Assuming the isotopic signature and concentration was the same 

in columns L2 and L1 at the end of the biostimulation, this lower apparent fractionation factor 

after suboxic elution suggests that the loss of sulfur, such as by oxidation and dissolution 

preferentially occurred for lighter isotopes. 

Acid volatile sulfide (AVS) content of the pre-column sediment was 0.3 μmol/g, or about 

0.1% of the total S content. The column L1 sediment intervals at the end of the biostimulation 

had AVS concentrations that ranged from 0.9 to 1.6 μmol/g, which accounts for less than 2% of 

the increase in total S (Table 3.8C). The column L2 intervals recovered after suboxic elution 

ranged from 0.2 to 1.2 μmol/g AVS, also a small percentage of the increase in total S. The lower 

AVS in column L2 suggests oxidation of part of the sulfide solid phases precipitated. The very 

low component of AVS suggests that the sulfide produced by sulfate reduction that was not 

transported out of the column had either transformed to FeS2 if precipitated with ferrous iron 

(Mauer and Rittmann, 2004) or as elemental S resulting from reaction of free sulfide with ferric 

iron solids resulting in oxidation of S-2 and reduction of ferric to ferrous iron (Li et al, 2009). 

These processes are illustrated in the following reactions:

FeS + H2S ↔ FeS2 + H2 

2FeOOH(s) + HS- + 5H+ ↔ 2Fe2+ + So + 4H2O.

Transformation of FeS to FeS2 by reaction with H2S is the primary pathway for pyrite 

formation under anaerobic conditions in groundwater systems (Mauer and Rittmann, 2004).
3-54

3.3.3.4 Total solid phase carbonate

The total carbonate content of sediments from intervals of column L1 (19.7±0.9 wt % as 

CaCO3), L2 (20.2±0.3 wt % as CaCO3) and L5 (20.6±0.7 wt % as CaCO3) were not significantly 

different than total carbonate measured in pre-column sediment (20.8±0.2 wt % as CaCO3) 

indicating that precipitation and dissolution during biostimulation and suboxic elution were small.

3.3.4 X-ray Spectroscopy and μXRF Mapping of Column Sediment

3.3.4.1 U oxidation state

XANES spectra of uranium in column L1, L2 and L5 sediments, and the pre-column ISR 

sediment as recovered from the field site, all are very similar to XANES spectra for a uraninite 

model compound with both the energy of the absorption maxima and near edge shoulder 

between 20960 and 20980 eV (Figure 3.14). In contrast, the U(VI) white line is at a higher 

0.0

0.1

0.2

0.3

0.4

0.5

0-2 2-4 4-6 6-8 8-10

Total S (mmol/g)

Distance from Inlet (cm)

L1 after reduction

L2 after suboxic elution

pre-column

-40.0

-39.0

-38.0

-37.0

0-2 2-4 4-6 6-8 8-10

δS34

Distance from Inlet (cm)

L1 after reduction

L2 after suboxic elution

pre-column

A

B

Figure 3.13 (A). Total sulfur concentration in column sediments by interval for column L1 

after biostimulated reduction and column L2 after suboxic elution, with pre-column sediment 

shown for comparison. (B). S isotopic composition of total sulfur expressed as parts per 

thousand δ34S (‰).
3-55

energy and the post edge shoulder is significantly more pronounced, as illustrated by a 

spectrum for andersonite U(VI) model (Figure 3.14). The oxidation state of uranium in 

sediments recovered from column L1 after biostimulated reduction with lactate and column L2 

recovered after suboxic elution was quantified by linear combination fitting of the background 

corrected and normalized XANES spectra. The best fits were obtained using a synthetic 

crystalline uraninite for the U(IV) component and andersonite (a uranyl carbonate mineral) for 

the U(VI) component. Best fits were determined by the smallest residual chi squared (χ2

) value 

of difference between the sample spectra and the linear combination fit of the components in 

the least squares fitting routine of SIXPACK software. The sum of the fractions of each 

component in the fits may not equal exactly one owing to uncertainty in the fit. Fitting of U 

XANES spectra typically can distinguish components to 5% at best (Singer et al, 2009), such 

that a component ≤ 5% cannot be detected. 

The U in all samples is dominated by U(IV) which comprises 85% or more of the total U in 

the sample (Table 3.9). The resulting fits are plotted along with sample spectra in Figure 3.14. 

Near equivalent fits to sample spectra were obtained using a single component uraninite U(IV)

model resulting in slightly larger χ2 values, and a U(IV) fraction of 0.95 to 1.01 (Table 3.9). 

These single component fits provide further evidence that U is predominantly in the U(IV) 

oxidation state. Column L1 interval R2 was measured both in December 2011 and July 2012. A 

small increase in the fraction of U(VI) was determined in later measurements suggesting that a 

small amount of oxidation occurred during storage. Samples were handled only in an anaerobic 

chamber, heat sealed in two layers of low permeability Mylar pouches containing oxygen 

scrubbers, and stored at -80°C. No measurable difference in U oxidation state was observed 

for column L5 sediment compared to the pre-column sediment. Although about 5% of the initial 

total U was transported out of column L5 during suboxic elution, the change in U speciation is 

not detectable by XANES. The eluted U was assumed to be the result of either desorption of 

residual U(VI) remaining after the ISR process or represents a readily oxidized fraction of U(IV). 

Overall, the U XANES indicate that the U uptake during biostimulation occurred primarily by 

reduction, and that U remaining after suboxic elution also was predominantly U(IV). In addition, 

little or no U(VI) produced during oxidation of bioreduced U(IV) is retained in the column by 

surface complexation to sediments with a majority of the re-oxidized U(VI) transported out of 

the column. This suggests the concentration of sites for U(VI) adsorption, such as on ferric 

oxide surfaces, is not sufficient to retain a significant fraction of the U(VI) by surface 

complexation. Limited U(VI) sorption during elution is consistent with a sharp increase in 

effluent U(VI) concentrations observed in the later part of suboxic elution at about 180 days (170 

PV) . 
3-56

Figure 3.14 U LII XANES spectra of column sediments after biostimulated reduction with lactate 

(L1) and suboxic elution (L2). ISR pre-column background sediment, and after suboxic elution of 

the precolumn sediment (L5) are shown for comparison. Numbers (e.g R1) refer to column 

subsection interval with lower numbers closer to inlet. Sample spectra are shown in solid lines 

and fits in open symbols. Models for crystalline U(IV) (uraninite, UO2) and U(VI) (andersonite), 

are shown for comparison.
3-57

Table 3.9 Summary of uranium oxidation state from best fits of bulk U XANES spectra of 

column sediment samples. U(IV) and U(VI) are expressed as fractions of total uranium from 

linear combination fitting. UO2 only represents fits to sample spectra using only crystalline 

uraninite model. UO2 + Andersonite represents fit results using crystalline uraninite as U(IV) 

component and andersonite for U(VI) component. * indicates measurement of second split of 

sample in July 2012, with first measurement in March 2010 for pre-column sediment, and 

December 2011 for L1-R2. All XANES were collected on U LII edge with the exception of the 

pre-column sample collected on the U LIII edge in March 2010. χ2 represents the sum of the 

square of difference between sample spectrum and the fit of the spectrum using the fraction of 

the components listed, and is specific to individual samples. Sample spectra and fits are shown 

in Figure 3.14.

Sample UO2 only UO2 + Andersonite

U(IV) χ2 U(IV) U(VI) χ2

L1-R1 1.05 3.30E-4 1.12 -0.08 2.70E-4

L1-R2 1.01 4.22E-4 0.96 0.07 3.81E-4

L1-R2* 0.97 5.63E-4 0.85 0.15 3.73E-4

L1-R3 1.01 4.93E-4 0.96 0.06 4.61E-4

L2-Ox1 0.99 6.82E-4 0.90 0.13 5.59E-4

L2-Ox3 0.99 8.85E-4 0.88 0.13 7.46E-4

L2-Ox4 0.98 8.23E-4 0.84 0.18 5.61E-4

L2-Ox5 1.01 6.99E-4 0.94 0.09 6.29E-4

L5-1 0.99 1.48E-3 0.85 0.17 1.24E-3

Precolumn* 0.95 1.19E-3 0.84 0.13 1.02E-3

Precolumn 0.99 1.20E-3 0.86 0.14 1.10E-3

3.3.4.2 Speciation of column sediment uranium

Extended X-Ray Absorption Fine Structure (EXAFS) spectra can provide information on the 

local bonding environment of an element through analysis of the spectrum and by comparison 

to model compounds. The EXAFS spectra for the ISR column L1 and L2 sediments recovered 

after biostimulation and suboxic elution are presented in Figure 3.15. The reciprocal k-space k3

-

weighted spectra of post biostimulated reduction column L1 samples are similar to spectra for 

nano-particulate and crystalline uraninite (UO2) between k of 2 and 6, for example the higher 

frequency oscillation that appears as a shoulder at about k of 4. The data quality above k of 6 

degrades rapidly with a weak similarity to lower frequency oscillations in the model spectrum. 

The spectrum for the pre-column sediment shares the low frequency features of the uraninite 

model compounds. The possible presence of higher frequency oscillations are masked by 

noise. The post suboxic samples from column L2 also appear to be intermediate to the nanopparticulate UO2 model and the pre-column sediment spectra. The comparison is more evident 

in the radial distribution function of the Fourier transform of the spectra (Figure 3.15B). All 

samples and the UO2 model have a distinct U-O shell at about 1.8 Å indicative of the oxygen 

coordination shell of U(IV) which dominates all of the samples. A second shell at about 3 Å is 

present in the pre-column sediment, but is of unknown identity. The shell is present when the k 

range of the spectrum used for the transformation is limited to k<7, below which the 3 Å feature 
3-58

appears as a shoulder on U-O peak indicating that the 3 Å peak is representative of a longer 

distance neighbor and is not the result of noise in the higher k range of the spectrum. The precolumn sediment does not have the 3.8 Å U-U shell (Schofield et al, 2008) indicating that the 

background U is a non-uraninite U(IV) species. The Kingsville Dome ore body contained 

uraninite in addition to several other forms including U in clays and associated with iron disulfide 

replacing detrital Fe and Ti oxides (Arrendondo, 1991). The lack of a 3.8 Å U-U shell in the precolumn sediment is consistent with oxidation and mobilization of uraninite U during the ISR 

process. This suggests that the remaining U is atomically dispersed in the sediment, perhaps 

sorbed or coprecipitated with other minerals in a form that is relatively recalcitrant to oxidation 

and remobilization. 

The 3.8 Å U-U shell is evident in the column L1 sample spectra and of similar magnitude to 

nano-partuclate UO2 but is diminished in magnitude from the crystalline uraninite model. The 

decrease in intensity of this feature relative to the U-O shell may be the result of small particle 

size and/or low crystallinity of the U(IV) precipitated during the biostimulated reduction. 

However, its presence in column sample spectra but not in the pre-column sediment is 

consistent with formation of a uraninite (UO2) like U(IV) precipitate during reduction. The 3.8 Å

shell also is present in the post suboxic elution samples from column L2, suggesting that U not 

remobilized during suboxic elution remains in this form and has not changed to another form. 

This observation is consistent with continued dominance of U(IV) in sediment after suboxic 

elution. The 3 Å shell observed in the pre-column sediment also is present in column samples 

but is of lower magnitude because of the contribution of the UO2-like U(IV)formed in the column 

to the spectra of column samples. 

Further analysis of the column sediment EXAFS spectra may provide information on the 

local bonding environment of U fitting of the spectrum. This process entails simulation of spectra 

through optimizing distance and coordination numbers of nearest and next nearest atoms in the 

element of interest’s coordination sphere to obtain best fit to sample spectrum. This approach 

requires knowledge of the identity and structure of likely components comprising the element in 

a sample. The contribution of other U species evident in pre-column sediment need to be 

accounted for or subtracted prior to fitting. Clearly, the limited quality of spectra, which in part is 

a function of the low U concentration, severely restricts the ability to derive unique fits. Thus, 

shell by shell fits have not been attempted with the current data.
3-59

Figure 3.15 (A) U LII k3

-weighted EXAFS spectra and (B) Fourier transformed radial distribution 

of EXAFS spectra (not corrected for phase shift) for ISR column sediments after biostimulated 

reduction with lactate (L1) and suboxic elution following biostimulated reduction (L2). Precolumn experiment sediment and the crystalline urananite model are shown for comparison. 

Letters and numbers (e.g. R1) refer to column subsection interval with lower numbers closer to 

inlet. Uranium concentrations listed are total concentrations measured by hot nitric acid 

dissolution and include pre-column background U.

3.3.4.3 Distribution of U on sediments

Micro-focused synchrotron X-ray fluorescence (μXRF) was used to map the distribution of 

uranium and other elements in thin sections of sediments recovered from the L1 and L2 column 

experiments and the pre-column sediments. This section describes the μXRF elemental maps 

of thin sections at 30 μm step size (SSRL beam line 10-2 “mesoprobe”), and at subsequent finer 

resolution (2 μm step size, SSRL beam line 2-3 microprobe). The distribution of U oxidation 

states was determined at selected locations using multi-energy imaging across the U LIII- edge, 

or XANES imaging. Multi-energy maps were also collected across the Fe K-edge as an indicator 

of the distribution of iron phases. U LIII-edge XANES spectra were collected at specific points 

(μXANES) both to verify the XANES imaging approach and for comparison to bulk XANES 

measurements.

A B

0

5

10

15

20

25

0 1 2 3 4 5 6

R (Å)

Fourier Transform Magnitude

Urananite

Crystalline UO2.0

L1-R1

230 μg/g

L1-R2

190 μg/g

L2-Ox1

170 μg/g

L2-Ox3

95 μg/g

Pre-Column

57 μg/g

Nano-particulate

Urananite

U-O U-U

-5

0

5

10

15

20

25

30

35

2 3 4 5 6 7 8 9 10

k (A-1)

χ(k)•k3
3-60

The pre-column sediment (57 μg/g total U) has localized areas of U (Figure 3.16). The 

sediment contains some larger grains of up to about 1 mm in diameter that are optically dark. 

These grains are primarily iron. U does not appear preferentially associated with large iron 

bearing grains in this thin section. Calcium is widely distributed through the thin section in 

smaller grains and likely includes a large contribution from the abundant calcite and to a lesser 

degree from feldspars. Two of the areas with higher U content were imaged at higher resolution 

(Figure 3.17). In panel A (Figure 3.17A), the μXRF map shows a single grain (~100 x 200 μm) 

containing U perhaps as a coating intermixed with Fe, along with other grains containing either 

Fe or Ca. The U content of this grain was not sufficient for μXANES data collection. The high U 

areas in panel B were of smaller size (<10 μm). The μXANES of one higher U spot (ISL Precolumn pt 1) indicates essentially all U(IV) (Table 3.10, Figure 3.18).

The content of U in the thin section of column L1 interval R1 (inlet end) was visually greater 

in the 30 μm beam size μXRF map (Figure 3.19) than the pre-column sediment. Distributions of 

Fe and Ca were similar to pre-column sediment. An enlarged view of region 2 shows the 

presence of U in a large Fe grain (~200 x 400 μm), along with smaller grains (<100 μm) with 

higher apparent U concentration (Figure 3.19C and D). Uranium oxidation state mapping of this 

large Fe grain (L1-R1B G2) and an adjacent high U grain showed U(VI) in localized areas with 

U(IV) distributed more diffusely within the iron grain (Figure 3.20A and B). On panel 2A of 

Figure 3.20, U μXANES at points within these grains show U is largely U(IV) (73 to 79% ) within 

grain L1-R1B G2 (pt 1 and pt 2), but is 57% U(IV) and 40%U(VI) in the smaller grain (Table 

3.10; Figure 3.18 L1-R1B 2 pt 1, 2, and 3). The Fe μXANES map of this area shows that the 

large grain is largely iron sulfide, with lower concentration areas of iron oxide adjacent to the 

grain (Figure 3.20C). Figure 3.21 shows Fe XANES spectra of individual points on grains that 

are consistent with reduced Fe, such as iron sulfide (L1-R1 g2 pt1) and magnetite (L1-R1 g2 pt 

2). The U oxidation state map of the smaller grains shown in Figure 3.20 indicate that U is up 

to two-thirds U(VI) (Figure 3.20D and E) and the U in these grains is intermixed with (D) or 

surrounded by Fe (E). Fit of the U μXANES taken at a point within grain 2C show a mixture of 

68% U(IV) and 34% U(VI) (Table 3.10; Figure 3.18 L1-R1B 2C). It is unknown if the U(VI) is 

part of the residual U remaining after the ISR xtraction of the ore body, or is the result of 

sorption of U(VI) during the column experiments. 
3-61

A

B

C

2

1

Figure 3.16 ISR pre-column sediment thin section: (A) Optical scan; (B) meso-scale 

XRF tri-color map of whole thin section panel at 30 x 30 um resolution with U shown 

in red, Fe in green, and Ca in blue. U is shown as difference between fluorescence 

above and below the U LIII edge to remove contribution from Rb fluorescence. (C) 

Area of interest for BL2-3 XRF mapping bounded by yellow. Scale bars are in units 

of microns.
3-62

. 

A B

Figure 3.17 X-ray microprobe XRF maps of region 1 and 2 for ISR pre-column sediment thin 

section shown in Figure 3.16. (A) μXRF map of region 1 at 5 x 5 μm resolution with iron shown 

in green, uranium in red and calcium in blue. Color brightness increases with concentration. 

(B) Region 2 scanned at 10 x 10 μm resolution at a single energy with the image shown for 

iron in green, total U in red and calcium in blue. Images are mirror of Figure 3.16C and rotated 

90° counter clockwise. Yellow circles mark grains where U LIII μXANES spectra were 

attempted. The spectrum for the circled area in panel B is shown in Figure 3.18 (ISR Precolumn pt 1). Scale bars are in units of microns.
3-63

L1-R2 T G1b

L1-R2 T G1 pt3

L1-R2 T G1 pt2

L1-R2 T G1 pt1

L1-R2 B G3

L1-R1 B G2c

L1-R1 B G2 pt 3

L1-R1 B G2 pt 2

L1-R1 B G2 pt 1

17160 17180 17200 17220

eV

U(IV) Model

U(VI) Model 

Absorption (Normalized)

L2-Ox1 B G1 pt 1

L2-Ox1 B G1 pt 2

L2-Ox1 B G1 pt 2b

L2-Ox3 B G1 pt 1

L2-Ox3 T G1 pt 1

ISR Pre-column pt 1

Figure 3.18 U LIII μXANES spectra of points on thin sections of column sediments after 

biostimulated reduction with lactate (L1), suboxic elution (L2) and ISR pre-column 

background sediment. Letter and numbers (e.g R1) refer to column, subsection, interval with 

lower numbers closer to inlet (e.g. Ox1), thin section panel (B =bottom; T = top), grain, and 

point on grain within each thin section. Sample spectra are shown in solid lines and fits in 

open symbols. Models for U(IV) (crystalline uraninite, UO2) and U(VI) (andersonite, uranyl 

carbonate) are shown for comparison. Sample IDs and linear combination fits of percentage 

U(IV) and U(VI) are shown in Table 3.10. Locations of point XANES data collection are 

shown in Figures 3.17, 3.20B, 3.20E, 3.23A, 3.24A, 3.24B, 3.26, 3.28C,and 3.30.
3-64

Table 3.10 Summary of uranium oxidation state from best fits of individual point μXANES 

spectra on column sediment thin sections. U(IV) and U(VI) are expressed as fraction of total 

uranium from linear combination fitting of crystalline uraninite (UO2) for the U(IV) component 

and either andersonite (AND) or U(VI) sorbed to ferrihydrite (FHY) for U(VI) components. U LIII

edge μXANES were collected using a 1-μm beam size at SSRL BL 2-3. χ2 represents the sum 

of the square of difference between sample spectrum and the fit of the spectrum using the 

fraction of the components listed, and is specific to individual samples. Sample spectra and fits 

are shown in Figure 20. Sample IDs represent column, interval section, thin section panel, grain 

and point on grain. Locations are shown on XRF maps of thin sections in Figures 3.17, 3.20, 

3.23, 3.24, 3.26, 3.28 and 3.30. * denotes U μXANES points on large iron sulfide grains.

Sample ID Components U(IV) U(VI) χ2

ISL Pre-column pt 1 UO2, AND 1.01 0.001 2.26E-3

L2-Ox1 B G1 pt 1 UO2, AND 0.71 0.32 1.17E-3

L2-Ox B G1 pt 2 UO2, AND 0.88 0.14 3.91E-3

L2-Ox1 B G1 pt 2b UO2, AND 0.85 0.27 4.33E-3

L2-Ox3 B G1 pt 1 UO2, AND 0.81 0.21 1.48E-3

L2-Ox3 T G1 pt 1 UO2, AND 0.88 0.13 4.54E-3

L1-R1 B G2 pt1 * UO2, AND 0.73 0.30 1.34E-3

L1-R1 B G2 pt2 * UO2, AND 0.79 0.30 1.96E-3

L1-R1 B G2 pt3 * UO2, FHY 0.57 0.42 1.45E-3

L1-R1 B G2c UO2, FHY 0.68 0.34 1.61E-3

L1-R2 B G3 UO2, FHY 0.60 0.43 7.89E-4

L1-R2 T G1 pt1 * UO2, AND 0.82 0.24 1.62E-3

L1-R2 T G1 pt2 * UO2, AND 0.91 0.12 1.62E-3

L1-R2 T G1 pt3 * UO2, AND 0.99 0.08 2.95E-3

L1-R2 T G1b UO2, FHY 0.69 0.33 1.65E-3
3-65

1 3 2

2

2

2

C

D

A

B

Figure 3.19 Column L1-R1 bottom panel of thin section: (A) Optical scan; (B) meso-scale XRF 

tri-color map of whole thin section panel with U shown in red, Fe in green, and Ca in blue, with 

areas of interest bounded by yellow. Scale bars are in units of microns. (C) is zoom of area 2 of 

meso-scale XRF map in B, with grains of interest denoted (e.g. 2A). (D) is a higher zoom of 

large iron grain (2A) in panel C bounded by yellow.
3-66

D E

B

A C

1

1

2

3

3

2

2

1

Figure 3.20 X-ray microprobe U oxidation state maps of grains 2A, 2B, and 2C in lactate 

column sample L1-R1 thin section shown in Figure 3.19. Panel A is a U oxidation state map of 

grain 2A where iron is shown in blue, U(VI) in green, and U(IV) in red. Panel (B) shows only 

U(IV) and U(VI). Total counts for small grain in upper left of 2A are 391 for U(IV) and 634 for 

U(VI). The total counts from the large grain are 5347 for U(IV) and 5962 for U(VI). Panel (C) is 

an iron μXANES map illustrating distribution of iron sulfide (red), iron oxide (green) and S 

(blue). These are mirror images of Figure 3.19 and rotated 90° counter clockwise. Points in 

panel B depict locations of U μ-XANES spectra shown in Figure 3.18 (L1-R1 B G2 pt 1, 2, and 

3), and in panel C locations of Fe μ-XANES spectra shown in Figure 21 (L1-R1 g2 pt 1 and 2). 

Panels (D) and (E) are U oxidation state maps of smaller grains (2B and 2C) shown in 

enlarged area of Figure 3.19C, with U(IV) in red, U(VI) in green, and Fe in blue. Total counts 

are 2151 for U(IV) and 3252 for U(VI) for the area imaged in D, and 5850 for U(IV) and 6918 

for U(VI) in E. The U μ-XANES spectrum collected at the center of grain 2C is shown in Figure 

3.18 (L1-R1 B G2c). Scale bars are in units of microns.
3-67

Figure 3.21 Fe XANES spectra of points on ISR column L1 sediment thin sections, and iron 

model reference mineral spectra. Points are shown in Figures 20C and 23B. 

Figure 3.22 is the 30-μm resolution XRF image of column L1 section R2 thin section 

showing similar points of elevated U separate from Fe or Ca bearing grains as observed in L1-

R1 (Figure 3.22C grain 1B), and a large Fe grain with U present on its exterior (Figure 3.22C

grain 1). The 2-μm resolution U μXANES oxidation state map of this grain (Figure 3.23A) 

shows that U is located primarily on the edges or surface of this iron grain and within voids or 

pores. Both U(IV) and U(VI) are intermixed at varying proportions depending on location. U 

μXANES at 3 points (Figure 3.23A) indicate 82 to 99% of the U is U(IV) (Table 3.10; Figure 

3.18) . This grain is mapped as iron sulfide by Fe μXANES imaging (Figure 3.23 B), with a 

diffuse area of iron oxide coating a large portion of the grain. Note that the brightness of the Fe 

oxide depicted in this figure is enhanced relative to iron sulfide since the Fe oxide color intensity 

is scaled to the maximum count rate for this component, which is about 20% of maximum FeS. 
3-68

The Fe oxide would be barely visible if the maximum was set equal to the FeS maximum. The 

Fe μXANES spectra for points 1 and 4 (on Figure 3.23B) are consistent with spectra for 

reduced iron phases such as FeS with spectra for points 2 and 3 similar to the ferrihydrite model 

compound (Figure 3.21). XANES spectra of siderite, vivianite, and carbonate green rust are 

distinctly different than spectra from the samples. 

The 2-μm resolution U μXANES image of grain 1b (Figure 3.24B) maps as predominantly 

U(VI) with some U(IV) (Figure 3.18, L1 R2 T G 1b). In contrast, the fit to the point XANES 

spectrum collected at the center of this grain is 70% U(IV) and 30% U(VI) (Figure 3.18). 

Additional point XANES spectra would be needed to resolve this. The high concentration U 

grain in the lower panel of the L1-R2 thin section (Figure 3.22E grain 3) was mapped with near 

equal proportions of U(IV) and U(VI) (Figure 3.24A). The point XANES spectrum near the 

center of this ~120 μm grain (L1 R2 B G3) yielded 60% U(IV) and 43% U(VI) (Table 3.10; 

Figure 3.18).

The 30-μm resolution XRF image (Figure 3.25B) of the thin section of sediment from 

column L2 section Ox1, which was recovered after suboxic elution, also is similar in U 

distribution to the column L1 thin sections. None of the large Fe grains had significant U 

concentrations. The high concentration U grains depicted in the enlarged area in Figure 3.25C

were imaged at 5-um resolution (Figure 3.26A), with grain 2B also imaged at 2-um resolution 

(Figure 3.26B). U μXANES indicate 88 and 85% U(IV) at two points on grain 2 (L2-Ox1 B pt 2 

and 2b) and 71% U(IV) at grain 1 (L2-Ox1 B pt 1; Table 3.10; Figure 3.18). These higher 

resolution maps indicated little Fe associated with these U(IV)-bearing grains.

The 30-um resolution XRF image of the top panel of sediment from column L2 section Ox3 

shows a lower abundance of U grains (Figure 3.27B) with no apparent U associated with large 

Fe grains. The higher resolution U μXANES oxidation state map of the high U grain (Figure

3.28C) show that this grain (L2-Ox3 T G1) is predominantly U(IV) consistent with U μXANES 

spectrum at the highest U concentration point within the grain that shows 88% U(IV) (Table 

3.10, Figure 3.18). The bottom panel (Figure 3.29B) of the L2-Ox3 thin section also shows a 

lower U abundance than L2-Ox1 in the 30-um resolution XRF image. A 2-um resolution U 

μXANES image of the grain (L2-Ox3 B G1) circled in Figure 3.29C was predominantly U(IV) 

that was both intermixed with and in separate zones from Fe(Figure 3.30). The U μXANES 

spectrum at the highest U concentration point in the upper part of the grain yielded 81% U(IV) 

(Table 3.10; Figure 3.18).
3-69

Figure 3.22 Column L1-R2 top panel of thin section: (A) Optical scan and (B) meso-scale XRF 

tri-color map with U shown in red, Fe in green, and Ca in blue. Area of interest is bounded by 

yellow and is enlarged in C. Bottom panel of thin section L1-R2: (D) optical scan, and (E) mesoscale XRF tri-color map of whole thin section panel with U shown in red, Fe in green, and Ca in 

blue. Area of interest (grain 3) is bounded by yellow. Scale bars are in units of microns.

A

B

1B

1

C

D

E

3
3-70

A

1

2

3 B

1

2

4

3

Figure 3.23 X-ray microprobe maps of thin section L1-R2 top panel, grain 1 of crescent shaped 

iron grain shown in Figure 3.22C, but rotated 90 counter clockwise, and mirror image. (A) is tricolor U oxidation state map U(IV) shown in red, U(VI) in green, and Fe in blue. Total counts in 

image area 12209 for U(IV) and 16374 for U(VI). (B) is iron phase map with iron sulfide shown 

in red, ferrihydrite (iron oxide) in green, and sulfur in blue. Sulfur and iron sulfide map together 

as purple. Points in A depict locations of U μ-XANES spectra shown in Figure 3.18 (L1-R2 T G1 

pt 1, 2, and 3), and in B locations of Fe μ-XANES spectra shown in Figure 3.21 (L1-R2 T g1 pt 

1, 2, 3, and 4). Scale bars are in units of microns.
3-71

A

B

Figure 3.24 (A) X-ray microprobe U oxidation state map of L1-R2 bottom panel grain 3 in 

Figure 3.22E with U(IV) shown in red, U(VI) in green, and Fe in blue. Circle denotes the 

location of U μXANES spectrum L1-R2 B G3 shown in Figure 3.18. Total counts in imaged 

area are 21746 for U(IV) and 28354 for U(VI). (B) U oxidation state map of L1-R2 top grain 

1B on Figure 3.22C with U(IV) in red, U(VI) in green, and Fe in blue. Circle denotes the 

location of the U μXANES spectrum L1-R2 T G1b. Total counts in imaged area are 694 for 

U(IV) and 3748 for U(VI). Scale bars are in units of microns.
3-72

A

B

C

Figure 3.25 Column L2-Ox1 bottom panel of thin section: (A) Optical scan, and (B) meso-scale 

XRF tri-color map of whole thin section panel with U shown in red, Fe in green, and Ca in blue. 

(C) shows area of interest for BL 2-3 microprobe mapping bounded by yellow in B. Scale bars 

are in units of microns.
3-73

1

2

A

B

2

2B

Figure 3.26 (A) X-ray microprobe XRF map of region thin section L2-Ox1 

shown in yellow rectangle in Figure 3.25C. Fe is in green, U red and Ca 

blue. Color brightness increases with concentration. Images are mirror of 

Figure 3.25 and rotated 90° counter clockwise. Grains 1 and 2 are 

bounded by yellow ovals. (B) Detailed XRF map of grain 2 at 5 x 5 um 

resolution. U μXANES locations are bounded by yellow circles. U μXANES 

spectra for these points are shown in Figure 3.18 and denoted L2-Ox1 B 

G1 pt 1 for grain 1 in A, and L2-Ox1 B pt 2 and 2b in B. Scale bars are in 

units of microns.
3-74

B

C

A

Figure 3.27 Column L2-Ox3 top panel of thin section: (A) Optical scan; (B) meso-scale XRF tricolor map of whole thin section panel with U shown in red, Fe in green, and Ca in blue. (C) is an 

enlargement of the area of interest for X-ray microprobe mapping bounded by yellow in B. Scale 

bars are in units of microns.
3-75

B

C

A

Figure 3.28 A) X-ray microprobe XRF map at 5 x 5 um resolution map of the area within 

yellow oval in Figure 3.27C thin section L2-Ox3 top, where U is shown in red, Fe in green, 

and Ca in blue. The map is a mirror image and rotated 90° counter clockwise of the area

in Figure 3.27. (B) 2 x 2 um resolution map of the high U grain circled in A showing U(IV) 

distribution in high area concentration increasing with color warmness (blue to red). (C) is

U oxidation state of the same area in B, where U(IV) is in red, U(VI) in green, and Ca in 

blue. Fe was not shown because of very low concentration relative to U(IV). U μ-XANES 

spectrum was collected at center of highest U area of the grain and shown in Figure 3.18 

(L2-Ox3 T G1 pt1). Total counts in imaged area are 2570 for U(IV) and 1049 for U(VI). 

Scale bars are in units of microns.
3-76

B

C

A

1

Figure 3.29 Column L2-Ox3 bottom panel of thin section: (A) Optical scan; (B) BL10-2 mesoscale XRF tri-color map of whole thin section panel with U shown in red, Fe in green, and Ca in 

blue. Area of interest for higher resolution μXANES imaging bounded by yellow is enlarged in C.

Scale bars are in units of microns.
3-77

Visual inspection of meso and microprobe XRF maps suggests that U is distributed widely 

over sediments in the thin sections, primarily on smaller grains (<100 μm), with U often in 

particles of only a few pixels in size (e.g see Figures 3.17B, 3.19, 3.25, 3.26 and3. 27). 

Integrating U counts over the whole sample area indicates higher U contents of postbiostimulated reduction and post-suboxic elution column sediments than pre-column sediment 

that parallels the trend in total concentration. This observation suggests that biostimulated 

reduction results in distribution of U over many sediment grains instead of being concentrated in 

localized zones. The finer grains appeared to have much lower Fe content than large grains 

(see below). U μXANES spectra show that a majority of U is U(IV) with a slightly greater 

component of U(IV) in the post-suboxic elution samples (Table 3.10). The point U μXANES 

measurements for small grains measured in column L1 thin sections ranged from 60 to 99% 

U(IV) (average 74%) and L2 ranged from 71 to 88% U(IV) (average 83%). By comparison bulk 

XANES of column L1 and L2 samples averaged 97and 89% U(IV), respectively, suggesting 

some oxidation may have occurred in the thin sections during sample processing and 

fabrication. In addition, only a limited number of these finer particles with sufficient zones of 

high U were located in the μXRF maps to allow for μXANES data collection. All of the grains 

analyzed had U distributed throughout the particles. It is unclear if the U in particles measured 

was in the sediment prior to biostimulation or precipitated during the column experiment. 

Coatings of U(IV) that may have formed on the finer grain sediments or as discrete particles 

during the column experiment were not observed in either the 2 or 30 μm spatial resolution XRF 

A B

Figure 3.30 X-ray microprobe U oxidation state maps of L2-Ox3 bottom panel grain 1, the 

circled grain in Figure 3.29C. (A) U oxidation state at 2 x 2 um resolution where U(IV) in 

red, iron is shown in green and Ca in blue. U(VI) is not shown because of very low 

concentration relative to U(IV). (B) U(IV) distribution of the same grain where

concentration increases with color warmness blue to red. U μ-XANES spectrum was 

collected at center of highest U area in top part of the grain and is shown in Figure 3.18 

(L2-Ox3 B G1 pt1). Total counts in XANES mapped area are 50078 for U(IV) and 2339 for 

U(VI).
3-78

maps. Instead, data collection focused on higher concentration U areas, likely biasing 

measurements towards larger grain size fractions. 

The sediment also contains some larger particles (>250 μm) that included grains comprised 

predominantly of iron sulfide minerals, some of which contain measurable U, either as coatings 

or within pores (see Figures 3.19, 3.20, 3.22 and 3.23). To test the contribution of the large 

iron grains to total U in a sample, the integrated total U, Ca, Fe counts in meso-probe XRF 

maps of the entire sample area of each thin section were compared with the integrated counts 

within 4 largest iron grains in each thin section. These grains ranged from 0.32 to 1.2 mm 

diameter and averaged 0.6 mm for 36 grains among nine thin section panels, and include grains 

L1-R1 B 2A (Figures 3.19 and 3.20), and L1-R2 T G1 (Figures 3.22 and 3.23). These large 

iron grains comprise 1.3% of the area of all thin sections but accounted for 12.5% of the Fe. In 

contrast, they accounted for only 1.8% of the total U counts indicating only a slight enrichment 

relative to the whole sample. Little difference was observed in this trend among the pre-column, 

post-reduction (L1) and post-suboxic elution sediments suggesting that the U present in these 

large grains is not preferentially precipitated or sorbed onto these large iron sulfide mineral 

grains during the biostimulated reduction, but that it is instead likely part of the U remaining after 

the ISR extraction. The U in the two large Fe grains measured ranged from 57 to 91% U(IV) 

(average 76%; Table 3.10), indicating a lower U(IV) component than measured in bulk samples, 

which may reflect either incomplete oxidation and leaching by the ISR process, or oxidation 

during thin section fabrication. No attempt was made to isolate sufficient mass of these large 

grains from the pre-column sediment for XANES data collection. Grain size separations and 

measurement of U content by gamma spectrometry indicate that >250 μm fraction comprised 

6% of mass of bulk sediment and contained 5% of total U in the pre-column sediment. The U 

distribution among grain size fractions was not measured after biostimulated reduction because 

of insufficient sample. Comparison of the integrated counts in coarse grains in pre- and postcolumn sediment thin section XRF maps suggest there was little change in U mass distribution 

during biostimulated reduction, and therefore do not contribute significantly to the total U 

measured by bulk XAS. Based on these observations, it is concluded that U in large iron grains 

likely is remnant uranium after ISR extraction of the ore body.

3.3.5 Microbial Assay

Biomass was very low in all ten aquifer sediment samples, which were recovered by drilling 

into the leached out ore bearing zone of the Kingsville ISR site. Because only sample 3712 B-C 

580-590 was retrieved as an intact core, and subsequently used for batch and column 

experiments, results for this material are presented. 

PCR products could not be obtained from this sample from the 30 cycle bacterial 16S rRNA 

gene in order to generate TRFLP profiles. However, the qPCR 40 cycle assays have higher 

sensitivity, and sulfate reducing bacteria (SRB) were detected by this method in all core 

samples. Geobacter was not detected in any of this core sample by qPCR. Cloning of the dsrB

gene PCR product revealed a high diversity of SRB at this site despite the low biomass, and 

representatives were from most families of SRB, but dominated by Desulfobulbaceae and 

Desulfobacteraceae. Closest sequences in the database tended to be uncultured organisms 

from contaminated environments such as oil field, petroleum contaminated sediment, aquifers 
3-79

impacted by landfill leachate, and a uranium mill tailings site, as well as estuarine and deep sea 

sediments (Table 3.11). Tentative genus identifications based on closest similarity to named 

organisms are Desulfotomaculum, Desulfosalina, Desulfofustis, Desulfobulbus, 

Desulfatibacillum, Syntrophobacteriaceae, Desulfosarcina, Desulfobacter, and Desulfonema.

Biomass was very low in the batch experiments sediments, even after amendment with 

electron donors for the 60 day duration of the experiment. As a result, no data were obtained 

from the H2 or acetate treatments. However, the lactate amended sample did contain Geobacter 

and SRB at low concentrations, 1.47 x 103 gene copies /gm SRB, and 2.6 x 102 gene copies/gm 

Geobacter. As this is the only Kingsville Dome sample with Geobacter detected, Geobacter

endpoint PCR was performed and followed by cloning and sequencing. Geobacteraceae 

sequences detected were most closely related to subsurface clade 1 of Geobacteraceae, as 

well as Pelobacter propionicus (Table 3.12).

Biomass also was too low in column sediments to provide sufficient DNA in extractions for 

microbial assay. Because of the small column size only 0.5 – 1 g of sediment was available for 

extraction. However, bacterial amplifications were obtained for TRFLP fingerprinting in the 2 

effluent filter samples, though not in the sediment column samples. Only faint amplification 

occurred with the column sediment DNA extractions, which were not sufficient for fingerprinting. 

One of these filters was taken during the Fe(III) reduction stage before sulfate reduction, as 

indicated by high dissolved iron and no measureable sulfide in effluent (days 40 to 60). The 

other was taken after onset of sulfate reduction evidenced by measurable effluent sulfide (days 

60 to 74). The TRFLP fingerprints were very simple with 2 main peaks, but the proportion of 

each peak is different in the samples. Bacterial 16S rRNA gene clone libraries were constructed 

for these 2 samples, and showed that during the Fe reduction phase, the dominant organism 

was Rhizobium sp., representing 76% of the sequences. Several sequences from the 

Clostridiales, including one most closely related to Desulfotomaculum acetoxidans, represented 

17% of the clone library, with several other minor groups (Table 3.13a).

Results for qPCR detected Geobacter only in the first effluent filter collected at the end of 

the Fe reduction phase, but with a low abundance of 5 x 103

 copies contained on the filter 

(Table 3.13). SRB also were detected at very low concentration also in this filter with 8.9 x 

102

copies contained on the filter. Community shifts are evident in the second effluent filter 

collected after the onset of sulfate reduction. Rhizobium sp. GN33-3 is still dominant, but the 

other Rhizobium (sp. 49) is now a minor part of the community. A Synorhizobium has increased 

and represents 22% of the clone library. Known sulfate reducers have become more abundant, 

with the sequence most closely related to Desulfotomaculum which only represented 2.4% of 

the sequences during Fe(III) reduction, now representing 12.2%. Also, a Desulfosporosinus sp. 

which was not detected in the first clone library now represents 7.3% of the clone library (Table 

3.13b). Results for qPCR detected Geobacter only in the filter taken during the Fe reduction 

phase, and the abundance was low, with 5 x 103 copies contained on the filter. SRB were 

detected at very low concentrations also in the Fe(III) reduction phase filter with 8.9 x 102 copies 

contained on the filter.

Sulfate reducers appear to be numerically dominant at Kingsville dome relative to 

Geobacter, and the diversity of SRB is high at Kingsville dome compared to Rifle,CO samples. 

Perhaps Geobacter are more sensitive to the ISR process (introduction of large quantities of 
3-80

oxygen to the ore zone) that this sample set had undergone, relative to SRB. If so, the 

remediation post ISR may rely more heavily on SRB than at a uranium mill tailings site. It is also 

possible that sulfate reduction is the more dominant terminal electron accepting process here 

for another reason. The Geobacteraceae types that were detected in the lactate treated 

microcosm, were dominated by the subsurface clade 1 that Holmes et al. (2007) found to be 

important in many subsurface environments. Sequences very closely related to Pelobacter 

propionicus in the D. acetoxidans clade were also abundant, though not as dominant as the 

subsurface clade1. This suggest that although Geobacter may be impacted by ISR processes 

or otherwise is less competitive with sulfate reducers, the Geobacteraceae that are there are 

consistent with those expected to have the ability to reduce U(VI). Although the total biomass is 

low in the Kingsville Dome samples, the diversity of sulfate reducers along with their numerical 

dominance over Geobacter imply that these organisms are important “microbial players” at the 

Kingsville dome site.

The column experiment had too low biomass to characterize bacteria in the small mass of

sediment available, but it was possible to characterize the bacteria from the filters taken at 2 

time points, at the end of Fe(III) reduction during the decline of effluent dissolved Fe, and after 

the onset of sulfate reduction. The dominant organisms were in the nitrogen fixing genus 

Rhizobium in both cases, but there is evidence of community shift favoring sulfate reducers 

consistent with the onset of sulfate reduction, as the sequence close to the sulfate reducing 

member of the family Clostridiales, Desulfotomaculum acetoxidans, increases in proportion, and 

a sequence close to the sulfate reducing Desulfosporosinus is also detected. 

Desulfosporosinus is known to be capable of reducing U(VI) (Suzuki et al. 2004). Members of 

the Clostridiales are also evident in both samples and the second most dominant group after 

Rhizobium. 

It is not surprising that Geobacter qPCR abundances were higher than SRB abundance in 

the effluent filter sample taken during Fe(III) reduction, and not detectable in the filter taken after 

the onset of sulfate reduction; however, it is surprising that given this result, that the clone 

library during Fe(III) reduction did contain SRB but no Geobacteraceae. Although only one SRB 

sequence in the Fe(III) reduction filter was detected, it is possible that if more clones had been 

sequenced, a Geobacteraceae may have been detected, better reflecting its importance at that 

time during the experiment. 
Table 3.11 Microbial assay of Kingsville Dome ISR pre-column sediment sample 3712 BC 580-59 showing microbi

abundance, and closest BLAST hits.

Group 

# 

Clones # 

Relative 

Abundance

% 

Closest Blast 

Hit

% 

Identity Source 

1 D06 1 1.3 AY337048.1 76 continental marg

sediment

2 E11 1 1.3 FJ748851.1 84 estuarine sedime

3 B07,C04 2 2.6 EF065047.1 89

landfill leachate 

groundwater 

4 G08,H03,F07 3 3.9 AY337204.1 94 continental marg

sediment 

5 A11,A12,A05, 3 3.9 FM212324.1 91

petroleum 

contaminated 

sediment

6 C11,A09,H06,H10,F10,D03,G01,G02,D08,F02,H11,F09 12 15.6 CP001147.1 77 Thermodesulfov

7 A08,D04 2 2.6 GU127914.1 80 Peatland 

8 A07,H07,E12 3 3.9 AY953397.1 84 mudflat estuary 

9 D02,G06,A03,F04 4 5.2 EF065071.1 80

landfill leachate 

groundwater 

10 G05,E10,C01,A04,D09,C10 6 7.8 HQ690093.1 90 Old Rifle aquifer 

11 D07,B05,B08,B04,B09,C08,C06,C07 8 10.4 HQ690093.1 89 Old Rifle aquifer 

12 B02,B10 2 2.6 AY953400.1 94 mudflat estuary 

13 C02,B03,B06 3 3.9 AY953397.1 84 mudflat estuary 

14 E08,E04,E07,H08,E01,F03,C09, 7 9.1 AY741574.1 85 estuary sedimen

15 A10 1 1.3 DQ250779.1 84

landfill leachate 

groundwater 

16 C05 1 1.3 AF244995.1 82 oligochaete wor

symbiont 

17 A02 1 1.3 AB263169.1 84 deep sea sedime

18 D05 1 1.3 EF065047.1 87 landfill leachate 

3-81
groundwater

19 F05,E09,G09,G03,F015 6.5 AF360646.1 93 desulfosarcina 

20 B12,A06,E03,D114 5.2 FR689594.1 89 steel corrosion 

21 C031 1.3 AB263171.1 92 deep sea sedime

22 C121 1.3 FR689565.1 92 steel corrosion 

23 F11,H09,F08 3 3.9 CP002364.1 87

Desulfobulbus 

propionicus

24 D121 1.3 FR689554.1 78 steel corrosion 

25 E02 1 1.3 EU725482.1 94 anaerobic granu

sludge bed 

3-82
3-83

Table 3.12 Geobacter clones from sediments recovered at the end of the lactate amendment 

batch uranium uptake experiment using the 3712 BC580-590 sediment.

Phylotype 

# 

Relative 

Abundance 

(%)

Closest 

Genbank relative 

(Blast)

Blast source description Sequence 

similarity %

1 12.5 HQ875511 Geobacter from paddy soil 95

2 12.5 EF668602.1 Riflegeoclone 24 100

3 25.0 X70954.1 Pelobacter propionicus 99

4 43.8 CP002479.1 Geobacter M18 100

5 6.3 EF668503.1 uncultured Geobacter 

M16clone

89

Table 3.13A Sequences obtained from effluent filter collected during later stages of Fe(III) 

reduction.

Phylotype 

# 

Relative 

abundance 

(%)

Closest 

Genbank 

relative (Blast)

Blast source description Sequence 

similarity 

(%)

Tentative Identity

1 45.2 GU994887.1 uranium contaminated 

groundwater

100 Rhizobium sp. GN33-3

2 31.0 HQ652582.1 magnetite drainage 

sample

100 Rhizobium sp. 

49(2011)

3 7.1 HQ133178.1 hexadecane degrading 

methanogenic 

consortium

98 Sedimentibacter sp. 

(Clostridiales)

4 7.1 FJ938130.1 agricultureal compost 

waste

94 Clostridia

5 2.4 CP001720.1 Desulfotomaculum 

acetoxidans 

95 Desulfotomaculum 

(Closridia)

6 2.4 EU331405.1 Anaeromyxobacter 

dehalogens

99 Anaeromyxobacter 

dehalogens

7 2.4 JN247803.1 Pseudoxanthomonas 

sp.

98 Pseudoxanthomonas 

sp.

8 2.4 HM057106.1 Pseudomonas sp.Y3-3 92 Pseudomonas sp.Y3-3
3-84

Table 3.13B Sequences obtained from effluent filter collected during onset of sulfate 

reduction. 

Phylotype 

# 

Relative 

abundance 

(%)

Closest 

Genbank 

relative (Blast)

Blast source description Sequence 

similarity 

(%)

Tentative Identity

1 48.8 GU994887.1 uranium contaminated 

groundwater

100 Rhizobium sp. 

GN33-3

2 22.0 AM084031.1 Synorhizobium sp. 100 Synorhizobium sp.

3 2.4 HQ652582.1 magnetite drainage 

sample

98 Rhizobium sp. 

49(2011)

4 7.3 NR_042202.1 Desulfosprosinus lacus 

STP12

99 Desulfosporosinus 

lacus STP12

5 2.4 FN689722.1 Sporotalea propionica 99 Sporotalea 

propionica

6 12.2 CP001720.1 Desulfotomaculum 

acetoxidans 

95 Desulfotomaculum 

(Clostridia)

7 2.4 HQ133178.1 hexadecane degrading 

methanogenic consortium

98 Sedimentibacter sp. 

(Clostridiales)

8 2.4 HQ384240.1 Acetobacterium wieringae 99 Acetobacterium 

wieringae

3.4 Discussion of Experimental Findings

3.4.1 Biogeochemical Processes During Biostimulation

The integrated column effluent concentrations for lactate, acetate, total dissolved CO2, 

sulfate, sulfide, iron, and uranium combined with “whole-column” changes in content of iron, S, 

and uranium (Table 3.5) are used in this section to provide insight on the biogeochemical 

processes occurring during the biostimulated reduction stage of the column experiments. Some 

of these processes are incorporated into model simulation of biostimulated reduction presented 

in Section 4.0, specifically sulfate, iron and uranium reduction. 

Complete consumption of lactate was occurring by day 18, with concurrent appearance of 

acetate in the effluent reaching a steady state concentration between 14 and 55 days of 3 to 3.5 

mM. Lactate consumption, presumably by oxidation to acetate, and subsequent oxidation of 

acetate by microbial processes requires equivalent reduction of other constituents to balance 

electron transfer from these electron donors. Likely terminal electron accepting processes 

(TEAP) under anaerobic conditions in aquifer sediments are sulfate, iron, manganese, and 

uranium reduction. An electron balance for electron donors (lactate and acetate oxidation) and 

TEAPs was constructed using the integrated column effluent data and solid phase analyses 

(Table 3.14). Oxidation of lactate to acetate produces 4 electrons per lactate molecule 

oxidized. Subsequent oxidation of acetate to CO2 produces 8 electrons, but could result in 

formation of formate as an intermediate step (Sharp et al, 2011), although no formate was 

detected during HPLC analysis of column effluent for lactate and acetate. Reduction of sulfate 

to sulfide (S-2) requires an 8 electron transfer, 1 electron for reduction of Fe(III) to Fe(II), 2 
3-85

electrons for U(VI) reduction to U(IV), and 2 electrons for Mn(IV) reduction to Mn(II). Total 

sulfate reduction was estimated from the sum of the integrated effluent dissolved sulfide and the 

increase in solid phase S, assuming that solid phase S increase is all S-2. This estimated 

amount of sulfate reduction requires 52.4 mmol of electrons, which is similar to the total 

electrons donated in integrated loss of lactate (49.6 mmol, Table 3.14) assuming all of the 

lactate consumption results in acetate production. The net consumption of acetate is defined as 

the production of acetate from lactate minus transport of acetate out of the column and equals 

5.94 mmol. Oxidation of this amount of acetate produces 47.5 mmol of electrons during the 

experiment if acetate is oxidized completely to CO2 and requires a TEAP other than sulfate, 

since sulfate reduction equals lactate oxidation.

Ferric iron reduction estimated from the integrated effluent iron and the increase in 1-hour 

0.5N HCl extractable Fe(II) from the sediment accounts for 0.47 mmol of electrons, with 0.51 

mmol of electrons if the 24-hour 0.5N HCl extractable Fe(II) is used instead to represent iron 

reduction. Both estimates account only for a small fraction of the electron balance for acetate 

oxidation (Table 3.14). Reduction of Fe(III) in phyllosilicate minerals by dissimilatory microbial 

iron reduction can be significant in aquifer sediments (Wu et al, 2012; Lee et al, 2012; Komlos et 

al, 2007). However, ferrous iron produced by reduction of Fe(III) silicates may not be 

quantitatively extracted by 0.5 N HCl since silicate phases likely are not dissolved by HCl. The 

iron concentration in hot HNO3 digest of column L1 sediments was on average 23% higher than 

iron in pre-column and column L5 sediments. Since HNO3 also does not dissolve silicates and 

since the influent did not contain significant dissolved Fe, the increase may be attributed to 

Fe(II) from reduction of Fe(III) in phyllosilicate minerals, if this Fe(II) component is soluble in hot 

nitric acid. If so, the estimated total iron reduction in column L1 increases to 1.95 mmol. 

Alternatively, the Fe(III)-silicate reduction can be estimated from the difference between Fe 

measured in hot HNO3 digest and in the total HF dissolution. The difference between HF and 

HNO3 iron was less for column L1 sediments after biostimulated reduction than in pre-column 

sediments. Assuming this change is the result of reduction of Fe(III)-silicate to a HNO3-soluble 

Fe(II) form, 1.0 mmol of Fe(III)-silicate reduction is estimated that did not result in dissolved Fe 

transport out of the column, for a total reduction of 1.16 mmol when including effluent Fe. Both 

of these estimates of Fe(III)-silicate reduction are insufficient to account for electrons produced 

during acetate oxidation (Table 3.14). In fact, if the total iron in column L1 determined by HF 

dissolution (14 mmol) was initially Fe(III) and was all reduced to Fe(II) during the biostimulation, 

the total Fe accounts at most for one third of total electron transfer needed to close the electron 

balance for acetate oxidation to CO2. 

Uranium and manganese reduction combined account for less than 1% of electrons 

produced by 97.1 mmol acetate oxidation (Table 3.14). The total uranium retained by the 

column (17.0 μmol), if assumed all is reduced to U(IV), accounts for 0.034 moles of electrons. 

The integrated effluent Mn was 7.7 μmol (0.015 mmol electrons). The total initial sediment Mn in 

column L1 was 0.40 mmol. If all Mn were reduced during biostimulation, it would result in only a 

small fraction of the electron balance (0.8 mmol or 1.8%).

Combined, the measured total sulfate, iron, uranium and manganese reduction at most 

account for 54.4 moles of the 97.1 mmol (58%) of the electrons produced during lactate and 

acetate oxidation. The deficit in the electron transfer mass balance indicates there must be 
3-86

either some other TEAP or transformation of acetate pathway that doesn’t consume electrons 

such as methanogenesis (Lovley and Klug, 1986). Dissolved oxygen entering the column 

cannot be ruled out completely but is insufficient to close the balance. Although the influent 

reservoir was continuously purged and the inlet tubing was jacketed with the anaerobic gas 

mixture, the peristaltic pump tubing was in the air. The 43 moles of electrons remaining in mass 

balance (Table 3.10) is equivalent to reduction of 10.7 mmol of oxygen. This would require a 6.6

mM influent dissolved oxygen concentration, which is more than 25 times saturation with 

respect to air. 

Since no other products of terminal electron accepting processes can be invoked or 

measured, fermentation of lacate to hydrogen, acetate to methane and/or biomass growth likely 

account for the remaining acetate consumption since electron transfer to a TEAP is not required 

(Lovley and Klug, 1986). However, the very low biomass present in post column sediments

limited DNA extractions making bioassay unfeasible. Therefore, biomass growth is unlikely to 

account for the remaining consumption of lactate and/or acetate. Alternatively, syntrophic 

growth of methanogens with sulfate reducing bacteria such as many of the Disulvofibrio and 

Desulfotomaclum species has been reported in which methanogens utilize the acetate produced 

by sulfate reducing bacteria (Muyzer and Stams, 2008). 
3-87

Table 3.14 Summary of electron transfer during biostimulated reduction in column L1. Lactate 

and acetate consumed are from integrated column effluent concentrations (see Table 3.5). 

Reduction of U is from integrated increase in sediment recovered from column L1. Sulfate 

reduction is the sum integrated dissolved sulfide in effluent and the increase in total sulfur in 

column sediments. Iron reduction is the sum integrated dissolved iron in the effluent and the 

extracted iron from column sediments (see footnotes). Manganese reduction is the sum of the 

integrated effluent dissolved Mn and the total sediment Mn, assuming that it is all reduced.

Constituent

Total during 

biostimulation

mmoles 

HCO3

produced

mmoles 

Number 

of 

electrons 

transfer

mmoles 

electrons

Electrons 

accepted from 

donor (%)

Electron donors:

C3H6O3

-

+ 2H2O → C2H3O2 + HCO3

-

+ 6H+ + 4eLactate → acetate 12.4 12.4 4 49.6 51

C2H3O2

-

+ 2H2O → 2CO2 + 7H+ + 8eAcetate → CO2 5.94 11.92 8 47.5 49

Total electrons from 

donors 97.1

Electron acceptors:

SO4

-2 + 8e-

→S-2 6.55 13.1 -8 -52.4 54

Fe(III) + e- →Fe(II) (a) 0.47 -1 -0.47 0.48

Fe(III) + e- →Fe(II) (b) 0.51 -1 -0.51 0.52

Fe(III) + e- →Fe(II) (c) 1.95 -1 -1.95 2

Fe(III) + e- →Fe(II) (d) 1.16 -1 -1.16 1.2

U(VI) + 2e- →U(IV) 0.017 -2 -0.034 0.04

Mn(IV) + 2e- →Mn(II) 0.008 -2 -0.016 0.016

Total TEAP -53 to 

-54.4 54.5 to 57.7

Unaccounted e- 43 42

Dissolved CO2 effluent 5.1

HCO3 produced (e) 24.3

Missing CO2 19.2

(a) Fe(II) sediment from increase in 1 hr 0.5 N HCl extraction compared to pre-column sediment plus 

effluent Fe

(b) Fe(II) sediment from increase in 24 hr 0.5 N HCl extraction plus effluent Fe

(c) Fe(II) sediment from increase in hot HNO3 extraction plus effluent Fe

(d) Fe(II) sediment from change in (HF total - hot HNO3 extraction) plus effluent Fe

(e) Assumes 1 mole CO2 produced during lactate oxidation to acetate; and 2 moles CO2 acetate to CO2
3-88

Since methane production was not anticipated it was not analyzed in the column effluent. In 

addition, because effluent sampling was not designed for methane collection, headspace 

measurement of sample bottles used would likely have been compromised by diffusional loss of 

methane through PTFE effluent tubing and loss during equilibration with effluent bottle 

headspace and subsequent venting during filling. Since methane was not measured, the extent 

of methane production was estimated by assuming that methane production results from 

consumption of the remaining acetate. In contrast to acetate, no methanogens have been 

described that utilize lactate (Muyzer and Stams, 2008). The near steady state acetate 

concentration of about 3.5 mM between 30 and 55 days (Figure 3.7B ), requires consumption 

of 7.5 mM acetate. Lactate is fully consumed during this period. Combined, iron reduction and 

sulfate reduction, in excess of lactate consumption, account for at most 10% of the 7.5 mM 

acetate. Thus, some other process is required to utilize the remaining 6.8 mM acetate. 

Fermentation of acetate produces 0.5 mole CO2 and 0.5 mole methane per mole of acetate. 

Therefore, if fermentation is assumed, this steady state acetate concentration would produce 

3.4 mM methane, a concentration that is above saturation for methane (2 mM, Martens and 

Klump, 1980) at which point bubble formation occurs. An equilibrium partial pressure of 2.3 

atmospheres in the column apparatus is calculated for this dissolved methane concentration 

using Henry’s law. The partial pressure resulting from production of methane to account for the 

remaining acetate is consistent with the back pressure increase observed during the 

biostimulation experiment (Figure 3.31A), which exceeded 25 PSI (1.7 atmospheres) at day 60, 

the limit of the pressure gage. Methane production has been shown to reduce hydraulic 

conductivity in sand columns as a resulting of induced pressure and bubble formation limiting 

flow (Sanchez de Lozada et al, 1994; Ye et al, 2009). Comparison of bromide tracer 

breakthrough for column L2 conducted after lactate was removed from the influent, with column 

L5 that did not undergo biostimulation shows faster initial tracer breakthrough in L2 but 

increased tailing (Figure 3.31B). Both are consistent with a decrease in permeability and 

preferential flow paths which can result either from gas pressure or biomass clogging (Baveye 

et al, 1998). Additionally, the back pressure in the column apparatus relaxed to the initial 

condition within three pore volumes after influent lactate ended. The rapid change suggests a 

decrease in gas pressure with cessation of production and flushing.
3-89

Figure 3.31 (A) Column L2 backpressure at influent end of column versus time. Pressure is in 

pounds per square inch above atmospheric pressure. (B) Bromide tracer versus volumes of 

tracer influent for column L2 at the end of biostimulated reduction, and for column L5 with no 

biostimulation to represent initial condition of columns, plotted as the ratio of effluent to influent 

bromide concentration. Bromide elution for column L2 is plotted as one minus the ratio of 

measured bromide to the initial influent concentration.

0

5

10

15

20

25

0 20 40 60 80

Pressure (PSI)

Time (Days; 1 Pore volume = 1.06 days)

Column L2 Inffluent Pressure

0

0.2

0.4

0.6

0.8

1

0 20 40 60 80

L2

L2 elution

L5

0

0.2

0.4

0.6

0.8

1

Bromide C/Co

Volume (mL)

Columns L2 and L5 Bromide Tracer

A

B
3-90

An alternative pathway to fermentation of acetate to methane is the production of H2 and 

acetate from lactate by acetogenic microbes, with H2 subsequently used by methanogens to 

produce methane by reduction of dissolved CO2. In this case, acetate is consumed by sulfate 

and iron reducing bacteria. The required methane and/or hydrogen production by these 

pathways is consistent with column back pressures and may account for consumption of 

remaining acetate (Muyzer and Stams, 2008). Based on changes in cell counts and presence 

of methanogens in biostimulation column experiments with Rifle aquifer sediments at low sulfate 

concentrations, Komlos et al (2008) invoked methanogenesis to account for consumption of 

acetate since SO4, Fe and U reduction accounted for only 1.5% of electrons from the acetate 

donor. However, because sulfate concentrations were high throughout the ISR column 

experiment, sulfate reducing microbes should outcompete methanogens for acetate and 

hydrogen (Oremland et al, 1982). As a result, methanogenesis would be limited based on the 

frequently observed sequence of sulfate reduction continuing until sulfate decreases to low 

levels prior to methanogenesis (Lovley and Klug, 1986) and the spatial separation of sulfate 

reduction from methanogenesis in sediments (Kuivila et al, 1989). However, since excess 

electron donor is present (continuous 10 mM lactate input) this hierarchy of electron accepting 

processes may not apply since the electron donor is not limiting (Komlos et al, 2008a). 

Unfortunately, the low concentrations of DNA prevented assay of microbial communities in 

column sediments after the experiments. Thus, further description of microbial processes to 

account for observed consumption of lactate and acetate during biostimulation is not possible. 

A net increase of 5.1 mmol total dissolved CO2 was estimated by integrating effluent sample 

concentrations (Table 3.5). Total consumption of lactate and acetate for the duration of the 

column experiments would result in 24.1 mmol of CO2 produced, 12.4 from lactate oxidation and 

11.9 mmol from acetate, assuming complete oxidation to CO2. Since there was no measurable 

increase in solid phase inorganic carbonate, a large fraction of CO2 by complete oxidation of 

lactate and acetate could not be accounted for. However, loss of CO2 exceeding the initial 15% 

pCO2 from outlet tubing by diffusion and from sample bottles during re-equilibration and 

headspace displacement as effluent bottles filled may have occurred. The amount of lactate 

and acetate utilized for cell growth resulting in an increased biomass is assumed to account for 

the increased rate of lactate consumption during the first 20 days of biostimulation, but cannot 

be calculated. 

3.4.2 Stability of Immobilized Uranium

The near complete removal of dissolved U(VI) from the influent over the course of the 

biostimulation stage with the resulting sequestered U primarily as U(IV) suggests that 

stimulation and growth of ambient microbial populations in the Kingsville Dome aquifer after ISR 

extraction by addition of electron donor and carbon source may be an effective means of 

removing dissolved uranium from groundwater following the ISR process. 

Previous studies have reported oxidative remobilization of U(IV) under anaerobic conditions 

resulting from continued reduction of Fe and Mn oxides which serve as terminal electron 

acceptors for U oxidation coupled with increased dissolved carbonate shifting the redox state to 

more favorable conditions for U oxidation (Wan et al, 2005). The remobilization of U was studied 

following biostimulated reduction in column experiments with aquifer sediments from the Old 
3-91

Rifle, CO, site inoculated with Geobacter metallireducens under low sulfate conditions. These 

experiments showed near 88% and 97% remobilization within 54 days when 0.27 mM dissolved 

oxygen (saturation with respect to air) or 1.6 mM nitrate were added to the influent, respectively 

(Moon et al, 2007). After 200 days of biostimulated reduction in similar column experiments 

under low sulfate conditions, re-oxidation of U(IV) resulted in the increase of effluent [U] to 

over 150 μM within 10 days of exposure to dissolved oxygen at saturation with air, with 60% of 

U remobilized after 60 days and nearly all by 120 days (Komlos et al, 2008b). The extent of U 

remobilization in these experiments suggested little adsorption of U(VI) by iron oxides or other 

surfaces on the column sediments and that oxidation of ferrous iron provided negligible 

buffering of U(IV) oxidation. Cell decay and ferrous iron oxidation accounted for 43% and 23% 

of oxygen consumption, respectively, with U accounting for 12% (Komlos et al, 2008b). In 

contrast, Sharp et al (2011) found little remobilization of biologically reduced U in columns with 

Rifle sediments with no added sulfate upon exposure to oxygen. Instead, the increase in the 

relative proportion of U(VI) in the middle and outflow end of the column was attributed to 

retention of U(VI) by sorption following re-oxidation of U(IV) near the inflow end of the column.

The presence of iron sulfides has been proposed to limit re-oxidation of U(IV) by serving as 

a redox buffer or armoring the U(IV) precipitate (Abdelouas et al, 1999). The timing and extent 

of U remobilization following prolonged biostimulated reduction of sulfate, iron and U in column 

experiments with inoculated Rifle sediments show significantly less U re-oxidation and a delay 

in release of U upon addition of oxygen or nitrate to column influent (Moon et al, 2009), 

compared to low sulfate experiments which had little or no iron sulfide precipitation. The iron 

sulfides formed during biostimulation served to partially limit biogenic U(IV) oxidation by 

competing for dissolved oxygen. Less than 1% of U was remobilized by 0.27 mM dissolved 

oxygen and no oxidation at 0.03 mM dissolved oxygen was observed through 50 days of 

elution. Sulfide oxidation accounted for 85% of oxygen consumption. Mackinawite (FeS) has 

been shown to inhibit UO2 oxidation provided that it is in physical proximity to the U(IV) phases 

and is in sufficient quantity to effectively scavenge oxygen to act as a redox buffer (Bi et al, 

2013). 

In the present study, the time dependence of the suboxic elution of U in column L2 provides 

a measure of the stability of the bioreduced U. Prior to introduction of dissolved oxygen to the 

column influent stream, effluent [U] averaged 0.01 μM over 30 pore volumes after U(VI) and 

lactate were removed from the influent. Because other processes, such as adsorption of U(VI) 

by sediment, can change (lower) the effluent dissolved [U], rates calculated from column 

effluent data are considered net rates. A net rate of U release of 9.8E-10 μmol/sec is calculated 

for the whole column by multiplying this near steady state [U] by the flow rate. Previous studies 

of remobilization or oxidation of reduced U report rates in terms of mol g-1 sec-1, where g is the 

mass of reduced (or total) uranium (in grams) in flow through or batch reactors (Bi et al, 2013; 

Sharp et al, 2011; Ulrich et al, 2009). Following this approach, dividing the net rate by the mass 

of U attenuated during biostimulation yields a remobilization rate of 2.4E-13 mol g-1 sec-1 which 

is four orders of magnitude slower than rates measured for synthetic and biogenic UO2 (Ulrich et 

al, 2009). The very low effluent [U] prior to introduction of dissolved oxygen combined with the 

predominance of U(IV) (>95%) in the sediment following biostimulation are consistent with 

negligible release of sorbed U(VI) from sediments, assuming there is no oxidation during this 

period. 
3-92

Pre-operational dissolved oxygen concentrations are not well known in groundwater of roll 

front U ore deposits targeted for ISR extraction, with most reported values below the lower end 

of the working range of dissolved oxygen meters (0.5 mg/L). A lower value of 0.2 mg/L (6 μM) 

was chosen to test suboxic conditions typically found in these deep aquifers. Column L2 was 

eluted with 19 pore volumes of AGW containing 6 μM dissolved oxygen (20 days), followed by 

62 PV at 12 μM (67 days). During this period a total of about 27% of the total U uptake during 

biostimulation was remobilized, likely the result of oxidation. The small increase in the fraction of 

U(VI) in column L2 sediments following suboxic elution (13% U(VI), see Table 3.9) compared to 

before introduction of oxygen (9% U(VI) for column L1) further suggests transport of [U] out of 

the column is largely the result of oxidation of U reduced during biostimulation. 

During the first 35 days of suboxic elution, effluent [U] increased linearly from 0.07 to 0.63 

μM (Figure 3.9). The corresponding instantaneous net U release rates increased from 3.5E-12 

mol g-1 sec-1 during the first 5 days following introduction of O2 to 2.5E-11 mol g-1 sec-1 after 35 

days. The rate of U release then began to increase as illustrated by the change in slope of 

effluent [U], with a constant rate of increase between 52 and 82 days after introduction of 

oxygen (Figure 3.9). During this second period of linear increase in [U] the net instantaneous 

rate of U remobilization increased from 1.7E-10 mol g-1 sec-1 after 55 days (52 PV) to 7.9E-10 

mol g-1 sec-1 by the end of the experiment. These rates are more than an order of magnitude 

slower than rates for synthetic abiotic and biogenic UO2 (Ulrich et al, 2009), for synthetic UO2 in 

the presence of mackinawite (Bi et al, 2013), and more than 4 orders magnitude slower than 

U(IV) release rates from biogenic UO2 produced during biostimulation of Rifle aquifer sediments 

(Sharp et al, 2011). Although the rates from those studies are from batch and flow through 

reactors, the rates calculated here from column data are likely more indicative of release rates 

expected in an aquifer following biostimulation since the column rates include reactive 

processes that would occur during transport in an aquifer. Overall, the rate of U remobilization 

increased by more than a factor of 200 during the 81 PV suboxic elution, with full mobilization of 

bioreduced U projected to occur after 150 PV. 

The observed increase in mobilization rate may be the result of depletion of other 

components that compete for dissolved oxygen. Oxygen consumption was attributed primarily 

to cell decay and iron oxidation in oxic elution in low sulfate biostimulation column experiments 

with U (Komlos et al, 2008b). In this study, oxidation of 0.11 mmol ferrous iron occurred based 

on the difference between whole column extractable ferrous iron before and after suboxic 

elution, represented by columns L1 and L2 (Table 3.5). Based on the cumulative effluent 

concentration of iron during suboxic elution, about 70% of the Fe release during the suboxic 

elution had occurred before the change in the rate of U remobilization occurred 35 days after 

introduction of suboxic AGW, with 84% total iron release at 52 days. The transport of iron out of 

column L2 during suboxic elution accounts for about 26% of the total decrease in sediment 

Fe(II) concentration measured (the difference between column L1 and L2). It is unclear if the reoxidized iron that remains in the column precipitates as iron oxides and provides sufficient sites 

for adsorption of remobilized U(VI) since there was only a small increase in U(VI) in the 

sediment following suboxic elution. However, Komlos et al (2007) reported that re-oxidation of 

bioreduced Fe silicates resulted in oxidized iron in silicates with similar structure to before 

reduction. This finding suggests that a large component of the re-oxidized iron remaining in the 

column likely is phyllosilicate Fe. Adsorption of U(VI) onto silicate minerals likely is minimal 
3-93

compared to iron oxides forming on sediment surfaces since no measurable loss of [U] was 

observed in columns without biostimulation activity (columns H1 and H2). 

Solid phase total sulfur decreased by 1.6 mmol in column L2 at the end of suboxic elution 

compared to column L1 (Table 3.8C). Very little change in effluent sulfate was measured during 

the suboxic elution indicating that sulfides did not oxidize completely to sulfate and/or the 

contribution of sulfate from sulfide oxidation was insufficient to result in a measurable change in 

effluent sulfate. In either case, net decrease in sediment sulfur after biostimulation (column L1) 

and suboxic elution (L2) had to result in net transport of dissolved S out of the column. Since 

total S measurements were only made at the end of each experimental stage no information on 

the rate or timing of sulfur oxidation during suboxic elution is available.

The estimated oxidation of Fe from the change in sediment Fe(II), 0.11 mmol, is less than 

10% of the decrease in total S, indicating a large fraction of sulfur reduced during biostimulation 

did not form iron sulfides. This is consistent with the negligible fraction of acid volatile sulfide 

measured in column L1 sediments at the end of the biostimulation (Table 3.8C). Thus, it is 

unclear which competing oxidation processes provided redox buffering prior to the change in U 

remobilization rate occurring after about 50 pore volumes in column L2. 

3.4.3 Comparison to Model Simulation of Biostimulated Reduction

The following chapter (4) presents simulations of column effluent during biostimulation 

based on the reactive transport models for biostimulated U reduction in groundwater (Yabusaki 

et al. 2011, Fang et al, 2009, Yabusaki et al, 2007). The model assigns sulfate reduction to 

occur primarily through lactate oxidation, which is reasonable since the electron transfer for the 

total sulfate reduction is equivalent to electron transfer required from oxidation of all lactate 

consumed to form acetate. For the modeled reduction to equal electron transfer from acetate 

oxidation, the model also assigns a pool of Fe(III) oxide and one pool of phyllosilicate Fe(III), 

along with minor contributions from U reduction and biomass growth. A second pool of 

phyllosilicate Fe(III) is invoked to account for lactate not consumed during sulfate reduction 

based on the change in effluent sulfate, which as discussed above, likely underestimates sulfate 

reduction. The total of the solid phase Fe(III) pools used in the model simulations, 240 

millimoles for the entire column (4.2 mmol/g times mass of sediment in the column, 57 g), 

greatly exceeds the 15 millimoles total iron present in each column initially based on HF 

dissolution of pre-column sediment (0.25 mol/g). This large iron pool was required in model 

simulation to balance electron transfer because no other microbial processes involving electron 

transfer or fermentation were included. Nonetheless, the simulations are able to capture the 

rate and extent of iron and uranium reduction in the column which may allow upscaling for 

prediction of biostimulated reduction at the field scale.

The modeling effort did not simulate the remobilization of U(VI) in response to the 

introduction of oxygen to suboxic levels. A more thorough understanding of mechanisms 

resulting in oxidation and remobilization of U reduced during biostimulation is required to 

develop reactive transport simulation for remobilization of bioreduced U(IV). Moreover, 

remobilization of U following bioremediation of a roll front ISR site will be a function of 

groundwater dissolved O2 concentrations. This will depend on the ambient dissolved O2
3-94

concentrations in the surrounding aquifer as well as on how effectively oxygen from the ISR 

process can be eliminated throughout the system, including low permeability zones that may 

bleed out dissolved O2 for long times. 

3.5 Summary of Experimental Findings

The laboratory column experiments with Kingsville Dome aquifer sediment recovered 

following ISR extraction under conditions geochemically relevant to the aquifer described in this 

chapter demonstrate the ability of the indigenous microbial population during biostimulation with 

lactate to reduce uranium resulting in precipitation of U(IV). The results provide a basis for 

evaluating the feasibility and design of field scale tests of biostimulated remediation of elevated 

groundwater uranium resulting from in situ leaching and recovery of uranium from sandstone 

uranium roll front ore deposits. The biostimulated reduction by lactate lowered dissolved 

uranium from 10 μM to <0.1 uM within 45 days. Near complete uptake of the influent 20 μM 

U(VI) continued through 82 days principally by reduction to U(IV). Slow release of U occurred in 

response to return to suboxic conditions with dissolved U increasing to 0.13 μM (<30 μg/L) over 

10 days (~10 PV) of elution with suboxic groundwater (6 μM dissolved oxygen). U 

remobilization occurred at a linearly increasing rate with effluent [U] increasing to 0.63 μM by 35 

days (33 PV). The rate and extent of U remobilization subsequently increased with [U] >13 μM 

after 87 days (81 PV) of suboxic elution at the end of the experiment. Analyses of sediments 

recovered from columns at end of the biostimulation and re-oxidation stages showed that the 

sequestered uranium was primarily U(IV) in the form of a uraninite-like phase. Significant 

findings are highlighted below.

1. The indigenous microbes in the Kingsville Dome sediment after ISR extraction were 

capable of reducing uranium in batch tests in the absence of dissolved oxygen when amended 

with electron donors. The timing and rate of reduction proceeded in order for the following 

electron donors: H2 > lactate > acetate. 

2. Hydrogen amended AGW (pH2= 0.05) was not sufficient to cause measurable attenuation 

of influent uranium in flow-through column experiments, likely because there was no added 

carbon source to promote biomass growth required for bioreduction of uranium, iron or sulfate.

3. Effluent concentrations from the two identical lactate amended columns (L1 and L2) were 

comparable throughout the biostimulated reduction stage. The summary below thus applies to 

both with differences noted.

4. Consumption of 10 mM influent lactate was observed starting at 5 days with effluent 

concentrations approaching 0 by 18 days. Complete consumption of lactate continued through 

to the end of the biostimulation period (82 days). Concomitant with lactate consumption, effluent 

acetate increased from 0 to 3 mM over the first 18 days and was relatively constant between 3 

and 3.5 mM through 55 days, suggesting steady state production and consumption of acetate, 

with the production rate from lactate oxidation exceeding acetate consumption. 

5. Effluent dissolved Fe increased starting at about 10 days indicating the onset of iron 

bioreduction with maximum concentrations exceeding 300 μM at 20-25 days, followed by 
3-95

decrease in effluent [Fe] to <10 μM by 45 days. The decrease in [Fe] was coincident with the 

appearance of measurable dissolved sulfide, indicating that the rate of sulfate reduction was 

producing sufficient dissolved sulfide to result in iron sulfide precipitation therefore limiting 

effluent Fe. A total of 0.16 mmol Fe was transported out of column L1 during biostimulation. An 

increase of 0.3 mmol of extractable Fe(II) was measured for column L1 with sediment Fe(II) 

concentrations increasing 4.8 to 7 μmol/g.

6. Dissolved sulfate in effluent decreased slightly from the influent concentration of 10.65

mM throughout biostimulation with an integrated loss of 1.4 mmol for column L1. Sediment total 

S increased by 0.06 to 0.2 mmol/g in column L1 intervals during biostimulation for a total 

retention of 4.9 mmol. The calculated decrease in δ34S isotopic signature (-40 to -43.2 ‰) of 

total S increase is consistent with microbial reduction of sulfate. Effluent dissolved sulfide 

increased to measurable concentrations after day 40, coincident with decrease in effluent 

dissolved Fe. Combined, effluent sulfide and solid phase S increase was 6.6 mmol. 

7. In anaerobic column experiments with ISR sediment the introduction of 10 mM lactate 

resulted in a decrease in the 20 μM influent dissolved U(VI) by 8 days (7.5 PV). Uptake of U 

occurred somewhat coincident to increase in effluent iron suggesting iron reduction may be the 

result of the same microbial process. Dissolved U decreased to <0.1 μM by 45 days (42 PV) 

with near complete uptake over the next 35 days that continued through the remainder of 

biostimulation. The cumulative uranium uptake of 16 μmol resulted in increasing the sediment 

U concentration by 0.3 μmol/g (70 μg/g) over the initial sediment U of 0.24 μmol/g. U uptake 

occurred preferentially near the column inlet with total U concentrations of up to 0.97 μmol/g.

8. Very low dissolved uranium (0.01 μM) was observed in the effluent for 30 PV after the 

electron donor and U(VI) were removed from the influent and prior to introduction of dissolved 

oxygen.

9. Upon introduction of 6 μM dissolved oxygen to the influent, effluent [U] increased linearly 

from 0.07 to 0.63 μM over 35 days (33 PV). The overall rate of U remobilization increased 

rapidly from this point with effluent [U] increasing to 5 μM by 55 days (52 PV) and was greater 

than 13 μM after the end of the experiment at 87 days (81 PV), with about 27% of bioreduced 

uranium being released. Effluent dissolved iron decreased somewhat linearly from 20 to 8 μM 

during the suboxic elution. No measurable difference was observed between influent and 

effluent sulfate throughout suboxic elution. 

10. Suboxic elution resulted in release of 2.1 μmoles of uranium, which is about 27% of the 

solid phase U uptake that occurred during biostimulation. Sediment Fe(II) decreased by about 

30% as a result of suboxic elution. Little change in reducible Fe(III) was measured over the 

course of biostimulation and suboxic elution compared to the pre-column sediment, suggesting 

the reduction and re-oxidation of iron occurred largely in phyllosilicate Fe(III)) minerals. Solid 

phase total S decreased 1.65 mmol (33%) during suboxic elution. 

11. XANES measurements showed that >95% of the U uptake occurred as the result of 

reduction and precipitation. EXAFS spectra of column sediments indicate that bioreduction of 

uranium may result in the formation of a uraninite phase. The low amount of sequestered 
3-96

uranium combined with the pre-existing U(IV) of a different form in pre-column sediments limited 

more conclusive structural determination of the U(IV) formed in the columns. Little change in 

the column sediment U(IV) fraction was measured after suboxic elution indicating that U(VI)was 

largely transported out of column as dissolved uranyl species with a negligible fraction of the reoxidized U(VI) adsorbed to sediments. 

12. Microfocused synchrotron XRF imaging showed that U is largely distributed among 

smaller grain sizes of sediment. Large iron sulfide grains were present and contain U which may 

be remnant U not leached during ISR process, and included both U(IV) and U(VI). These grains 

comprised a small fraction of the total U.

13. Microbial assay of ISR sediment used for the column experiment showed that despite 

low biomass there was a high diversity of sulfate reducing bacteria (SRB) in the sediment 

recovered from the aquifer that included representatives from most families of SRB, but 

dominated by Desulfobulbaceae and Desulfobacteraceae. Geobacter was not detected, 

suggesting an important role of SRB for reduction of iron and uranium, as well as sulfate, as 

previously observed (Lovley et al, 1993; Beyenal et al, 2004). Bioassay of an effluent filter 

collected on the falling limb of the dissolved Fe peak show presence at low abundance of both 

Geobacter and sulfate reducing bacteria during iron reduction. Another effluent filter, collected 

during sulfate reduction, showed a community shift toward increasing SRB and a concomitant 

decrease in the iron reducing Geobacter. Biomass in sediments recovered from columns was 

too low to provide sufficient DNA for community identification.

14. Geochemical mass balance for the biostimulation period indicates other processes are 

required to account for 42% of electron transfer resulting from the observed consumption of 

lactate and acetate in excess of sulfate, iron and uranium terminal electron accepting 

processes.

15. Suboxic elution of Kingsville ISR sediment that had not been subjected to biostimulated 

reduction showed that 4.6% of the sediment-associated U was mobilized after 44 pore volumes, 

which included more than 30 pore volumes of low oxygen (~6-12 μM) influent solution. 
4-1 

4 Simulation of Uranium Bioremediation in ISR Sediment

Columns

We describe the modeling of transport and uranium biogeochemistry during biostimulated 

reduction in a laboratory column experiment that was performed by the USGS using the postISR sediments from Kingsville Dome (Chapter 3). Recent investigations in shallow uraniumcontaminated aquifer systems (Anderson et al., 2003; Vrionis et al., 2005; Williams et al., 2011; 

Wu et al., 2006a; Wu et al., 2006b) have demonstrated that the stimulation of indigenous 

dissimilatory metal-reducing bacteria, via the addition of an electron donor (i.e., ethanol, lactate, 

acetate), can reduce soluble hexavalent uranium [U(VI)] to immobile solid-phase U(IV) 

(Anderson et al., 2003; Vrionis et al., 2005; Williams et al., 2011). While the bioremediation 

principle is common to both the shallow uranium groundwater plumes and the relatively deep 

uranium ISR aquifers, there is an important conceptual difference. In the shallow sites, the 

conditions favorable for the bioreduction of U(VI) need to be artificially maintained; whereas at

the post-ISR sites, the intent of the engineered biostimulation is considered a temporary action 

that is designed to return the geochemical conditions to the naturally low redox potential 

associated with the uranium roll front deposit. 

We use the systematic and quantitative coupling of process models to describe the interplay 

between the saturated flow, microbiology, and geochemistry during lactate biostimulation of the 

Kingsville Dome sediment column experiments. Modeling of these experiments is used to 

understand and quantify uranium behavior in the context of 1) biostimulation with lactate 

electron donor, 2) functional microbial groups representing Fe(III)-reducing bacteria (FeRB), 

which are also responsible for the bioreduction of U(VI), and sulfate-reducing bacteria (SRB) 

responding to the biostimulation, and 3) geochemical response to the biostimulation products. 

While both columns, L1 and L2 in the laboratory study described in Chapter 3 experienced 

83 days of lactate biostimulation, Column L2 was subsequently used to investigate non-reactive 

tracer transport and the impact of suboxic conditions on uranium behavior following 

biostimulated reduction. In this chapter, the focus of the column experiment modeling is on the 

engineered uranium bioreduction and immobilization in Column L1. The effluent profiles of 

both columns were very similar throughout the biostimulated reduction stage.

4.1 Uranium Bioremediation Modeling Studies

A particularly relevant reactive transport modeling study is NUREG/CR-6870 (Davis and 

Curtis, 2007), which described the use of the PHREEQC computer code (Parkhurst and Appelo, 

1999) to analyze the geochemistry during groundwater restoration at the Ruth ISL (Schmidt, 

1987) in Wyoming. In this case, models that consider groundwater flow, solute transport, and 

geochemical reactions are used to estimate the number of pore volumes and therefore costs 

associated with groundwater restoration. One finding was that slow desorption from the 

sediments resulted in the persistence of U, Se, and V above baseline concentrations despite 

tens of pore volumes of groundwater flushing. 
4-2 

Saturated flow and multicomponent biogeochemical reactive transport modeling studies 

were also performed to simulate uranium bioremediation in shallow contaminated groundwater 

plumes. These include the 2002, 2003, 2007 and 2008 field biostimulation experiments at the 

Rifle IFRC site (Fang et al., 2009; Li et al., 2010; Li et al., 2009; Yabusaki et al., 2007; Yabusaki 

et al., 2011). A comprehensive reaction network was developed from these studies that 

included biologically-mediated TEAPs for solid phase Fe(III), aqueous U(VI), and aqueous 

sulfate; aqueous and (nonelectrostatic) surface complexation for Fe(II) and U(VI); calcite, 

siderite, FeS, S secondary minerals; sulfide promoted dissolution of Fe(III) minerals, cation 

exchange, protonation/deprotonation of Fe(III) surfaces, and major ion chemistry. In this case, 

the integration of the abiotic chemistry with the TEAPs in the modeling is necessary to account 

for the impact of biostimulation products on uranium behavior. In laboratory and field 

experiments, the bulk of the bicarbonate and reduced phases (e.g., Fe(II), sulfide) produced by 

the biologically-mediated reactions are not observed in solution. This underscores the 

importance of accurately representing the solid phase reactions (e.g., mineral reactions for 

carbonates, iron, and sulfur; iron and uranium surface complexation) that control the pH, Eh, 

alkalinity, and aqueous components of interest. 

An important challenge identified in these uranium bioremediation modeling studies (Fang et 

al., 2009; Williams et al., 2011) is the impact of biologically-mediated reaction products (e.g., 

biomass, bicarbonate, Fe(II), U(IV), H2S) on the behavior of uranium (Englert et al., 2009; Li et 

al., 2010; Li et al., 2009). In particular, the sensitivity of uranium mobility to pH, Eh, alkalinity, 

calcium, and reactive surface area requires detailed biogeochemical process modeling to 

mechanistically simulate changes to the aqueous and solid phase chemistry induced by acetate 

biostimulation (Davis et al., 2006; Dong et al., 2005; Li et al., 2009; Liu et al., 2005; Luo et al., 

2007). Biologically-mediated oxidation of acetate produces a large amount of bicarbonate that 

directly affects the alkalinity, and indirectly affects the pH, calcium and reactive surface area 

through precipitation of carbonate minerals (Li et al., 2009). The reductive dissolution of Fe(III) 

minerals and the precipitation of sulfides and elemental sulfur affects pH and reactive surface 

area (Poulton et al., 2004). CaUO2(CO3)3

2- and Ca2UO2(CO3)30

(aq) complexes (Dong and 

Brooks, 2006), which are predicted to be predominant for the Kingsville Dome post-ISR

geochemical conditions, have been shown to affect U(VI) bioreduction (Brooks et al., 2003) and 

adsorption (Zheng et al., 2003). Thus, the inclusion of a reaction network that accounts for 

uranium, terminal electron-accepting processes (TEAPs), biomass, major ion chemistry (Ca, 

Mg, K, Na, carbonate, sulfate, Cl), reduction products [Fe(II), U(IV), H2S], aqueous complexation 

(e.g., Ca-UO2-CO3), surface complexation [Fe(II), U(VI)], and precipitation/dissolution (goethite, 

calcite, FeS, S, UO2) is critical to capturing important biogeochemical interactions during 

biorestoration of a uranium ISR site. 

4.2 Findings from Column Experiments using Kingsville Dome ISR 

Sediments

The column experiments in Chapter 3 were designed to investigate the potential for 

stimulating indigenous microorganisms in post-ISR Kingsville Dome aquifer sediments. The 

findings from the biostimulation stage of column L1 were the following:
4-3 

• After 14 days, the column effluent indicated that the continuous influx of 10 mM lactate 

was completely consumed during transport through the column (Figure 4.1). At this 

point in the experiment, the nonreactive transport travel time through the length of the 

column was 1.1 days. 

• Acetate, produced by the lactate oxidation reaction(s), increases from 0 and plateaus at 

3 mM at day 14 in the column effluent. From day 55 to 83 (end of the bioreduction 

phase of the experiment), the effluent acetate concentration increases to ~10 mM. This 

means that before the end of the experiment, nearly all the acetate produced by lactate 

oxidation leaves the column unreacted.

• Fe(II) increases to a peak of ~320 μM after 30 days, after which it decreases to less than 

10 μM at day 45 (Figure 4.2). There are two minor peaks at 4 and 15 days. 

• Continuous influx of 20 μM U(VI) was transported through the column with negligible 

attenuation until the lactate was being completely consumed at day 14. After day 14, 

aqueous U(VI) in the column effluent began to decrease. The rate of U(VI) reaction 

progressively increased until day 40 when essentially none of the continuously influent 

20 μM U(VI) was found in the column effluent (Figure 4.3), indicating complete 

attenuation of dissolved U. 

• Sulfate consumption during transport through the column was ~1 mM from the 

continuous influx of 10.65 mM sulfate (Figure 4.4). The sulfate consumed in the column 

was relatively small despite influent lactate and produced acetate concentrations that 

were ~10 mM. 

Over the first 14 days of the column experiments, the rate of microbially-mediated oxidation 

of the continuously influent 10 mM lactate progressively increases. After 14 days, lactate 

transported through the column is completely consumed. At this point, 31 and 35 pore volumes 

of influent solution have passed through columns L1 and L2, respectively. Assuming one mole 

of acetate is produced for each mole of lactate reacted with no other microbial process 

consuming lactate, 10 mM acetate is being generated in the column at this time. From day 14 

to day 55, effluent acetate is ~3 mM, meaning that 85% of the electron donors available to the 

column were consumed (i.e., 10 mM lactate and 7 mM acetate). From that point in time to the 

end of the experiment at 83 days, effluent acetate increases and approaches 10 mM, while 

lacate is fully consumed. This implies that at the end of the experiment, nearly all of the acetate 

produced is unreacted. 
4-4 

Figure 4.1 Effluent lactate (blue symbols) and acetate (red symbols) concentrations from 

column 1 (solid symbols) and column 2 (open symbols).

The increase in aqueous Fe(II) during the first 30 days of biostimulation followed by a 

decline to near zero is similar to observations in shallow aquifer sediments. The increase is 

consistent with the initial activity of the Fe(III) TEAP reactions, while the decrease from day 30 

to 45 is consistent with the production of sulfide via sulfate bioreduction that reacts with the 

aqueous Fe(II) to form FeS, effectively titrating Fe(II) from solution. 
4-5 

Figure 4.2 Effluent Fe(II) concentrations from column 1 (blue diamonds) and column 2 (red 

squares).

Other behaviors exhibited by the Kingsville sediment column experiments are not similar to 

observations in biostimulation experiments with shallow aquifer sediments. For the first 10 

days, effluent uranium is generally at influent concentrations as opposed to the rapid uranium 

bioreduction exhibited for acetate biostimulation in shallow aquifers. Only until half or more of 

the influent lactate is consumed (> 10 days) is a measurable decrease in effluent U(VI) 

concentration observed. At day 20, ~2.5 mM of the 20 mM influent U(VI) has been removed. 

By Day 40, nearly all the influent U(VI) has been removed. Thus, the most effective U(VI) 

removal occurred after the point in time when the influent lactate was being fully consumed

during the 1.1 day transit of column L1. 
4-6 

Figure 4.3 Effluent U(VI) concentrations from column 1(blue diamonds) and column 2 (red 

squares).

In the shallow aquifer system, extensive sulfate bioreduction appears to interfere with the 

efficiency of U removal through SRB competition for electron donor and the production of 

bicarbonate which favors desorption of U. While the sulfate concentrations, ~10 mM, are similar 

to the Rifle shallow aquifer, only 10 to 20% is removed during the experiment, putatively by 

bioreduction to sulfide. While the sulfide generated is still more than sufficient to completely 

react the ~300 μM Fe(II) generated by Fe(III) bioreduction, the persistence of sulfate in the 

presence of lactate and excess acetate is unlike the Rifle biostimulation behavior. 
4-7 

4.3 Conceptual Biorestoration Model

4.3 Conceptual Bioresoration Model

Based on these observations, a conceptual model of processes controlling the Fe(III), U(VI), 

and sulfate TEAPs during lactate biostimulation in the column experiments was postulated. A 

preliminary assumption is that the lactate and acetate electron donors are principally consumed 

in TEAP reactions involving Fe(III), sulfate, and U(VI). 

• Sulfate bioreduction occurs principally through lactate oxidation. This is consistent 

with the situation at the end of the column experiment when the removal of 1-2 mM 

sulfate occurs while acetate is largely unreacted. Based on the stoichiometry of the 

sulfate TEAP reaction with biomass in Fang et al. (2009), this means that 3 to 6 mM 

lactate is being consumed. 

• Two pools of phyllosilicate Fe(III) are used in the model, one pool that is bioavailable 

for the lactate reaction, and one pool that is bioavailable for the acetate reaction. As 

only a relatively small amount of lactate is consumed for sulfate bioreduction, the 

only other TEAP capable of continuously consuming the remainder of the 10 mM 

influent lactate is through the Fe(III) terminal electron acceptor, which motivated the 

use of an Fe(III) TEAP reaction that was exclusive to lactate.

Assuming stoichiometric oxidation of lactate to acetate, the absence of lactate in the 

column effluent after 14 days means that the equivalent of 10 mM acetate is 

continuously being generated by the lactate oxidation. From day 14 to day 55, 

effluent acetate concentrations are fairly constant at ~3 mM, after which it increases

to ~10 mM by the end of the 83 day experiment. That means that ~7 mM acetate is 

consumed continuously between 15 and 55 days. Once again, the Fe(III) TEAP was 

Figure 4.4 Effluent sulfate concentrations from column 1(blue diamonds) and 

column 2 (red squares).
4-8 

the most straightforward process capable of continuously consuming ~7 mM acetate. 

The increase in effluent acetate concentrations to ~10 mM after day 55, however, is 

consistent with the depletion of the Fe(III) pool. Since Fe(III) must not be limiting to 

account for the continuous oxidation of the influent lactate, the assumed pool of 

Fe(III) that is bioavailable for acetate-based reaction must be finite. 

• Most of the biogenic Fe(II) is solid-associated. The effluent aqueous Fe(II) 

concentrations, which peak at ~300 μM after ~30 days represent a small fraction of 

the biogenic iron. In the Fe(III) TEAP reactions, 4 and 8 moles of Fe(II) are produced

for every mole of lactate and acetate oxidized, respectively. Thus, nearly all of the 

Fe(II) produced through TEAP reactions is not in the effluent. While there are 

multiple pathways for Fe(II) to be associated with the solid phases, the bioreduction 

of phyllosilicate Fe(III) (iron-containing clays) in which Fe(III) is reduced in place 

without the release of aqueous Fe(II) is the simplest. Phyllosilicate Fe(III) has also 

been shown to be more bioavailable than the crystalline Fe(III) pool (Komlos et al., 

2008b). 

• Uranium and goethite bioreduction are acetate-based and inhibited by lactate. 

Unlike the relatively rapid increase in effluent Fe(II) concentrations, U(VI) 

concentrations do not decrease until lactate is almost completely consumed. This 

lag in the onset of uranium bioreduction is consistent with lactate inhibition. In the 

U(VI) TEAP reaction with biomass from Fang et al. (2009), one mole of acetate can 

bioreduce three moles of U(VI). Thus, only ~7 μM acetate is needed to catalyze the 

conversion of the 20 μM influent U(VI) concentrations . This is negligible and 

probably impossible to discern in comparison to the millimolar acetate 

concentrations. 

The reductive dissolution of bioavailable crystalline Fe(III), represented by goethite in 

the model, is assumed to be the largest source of dissolved Fe(II) in the column. Its 

contribution to the effluent Fe(II) peak at ~30 days is also consistent with a lagged 

response from lactate inhibition. As identified in Fang et al. (2009), the amount of 

goethite reduced to Fe(II) is much greater than the observed aqueous Fe(II) in 

solution. This means that most of the Fe(II) produced by the acetate-promoted 

goethite reduction becomes solid associated through adsorption and secondary 

mineral formation (e.g., FeS, FeCO3). 

• The progressive increase in the uranium bioreduction rate over time is consistent 

with a biomass dependent rate. From day 10 to day 40, the progressive U(VI) 

decrease in the column effluent is consistent with a bioreduction rate that is 

dependent on the biomass, which is increasing with continuous acetate stimulation. 

This requires the direct incorporation of biomass into the reaction rate law, i.e., 

augmenting the dual monod approach with a biomass term. 

• The specification of initial Fe(III) concentrations is critically important to match the 

timing of the effluent acetate behavior. The relatively rapid increase in effluent 

acetate concentrations to ~8-10 mM after day 55 is consistent with the depletion of 
4-9 

the crystalline and phyllosilicate Fe(III) pools that are bioavailable to the acetatepromoted bioreduction reactions. 

We consider this conceptual model a preliminary attempt that should be updated as more 

details of the biologically-mediated processes become available. While the postulated 

processes are based on our experience in biostimulation field experiments in shallow aquifers, it 

should be noted that the hydrobiogeochemistry exhibited in these column experiments using 

post-ISR sediments from the Kingsville Dome uranium deposit is distinctly different (e.g., lag in 

U(VI) bioreduction, low sulfate bioreduction in the presence of excess acetate). Thus, 

alternative conceptual models for the column behaviors are possible. 

4.4 Simulator Description

In order to model how the microbial community will respond in subsurface environments 

undergoing bioremediation, the microbially-mediated reactions are specified in a reactive 

transport model that can define the spatial and temporal variability in the biogeochemical 

environments. We used the HYDROGEOCHEM code (Yeh et al., 2004) to resolve the detailed 

coupled processes and model the large number of reactive components in the column 

experiments with the Kingsville Dome post-ISR sediments. HYDROGEOCHEM is a 

comprehensive simulator of coupled fluid flow, hydrologic transport, heat transfer and 

biogeochemical reactions under variably saturated conditions in two or three dimensions. It uses 

a finite element method based on unstructured grids, iteratively solving fluid flow, heat transfer 

and reactive chemical transport equations. The numerical model incorporates advection, 

diffusion with Fick’s law modified for the inclusion of dispersion, and mixed kinetic and 

equilibrium reactions. An operator splitting solution method is used for flow, transport, and 

reactions, each of which employs implicit time-stepping schemes. A general reaction-based 

approach to biogeochemical processes was implemented in the code so that it can be applied to 

a wide range of problems (Fang et al., 2003; Fang et al., 2006). Programmed reaction types 

include aqueous complexation, adsorption-desorption, ion-exchange, oxidation-reduction, 

precipitation / dissolution and microbially mediated reactions. 

HYDROGEOCHEM provided the framework for incorporating and coupling the flow and 

reactive transport process models for this column study. In particular, the multicomponent 

biogeochemical solver technology (Fang et al., 2006; Fang et al., 2003) addressed the 

complexity of mixed kinetic and equilibrium reactions that described the aqueous and surface 

complexation, redox, mineral precipitation and dissolution, and microbially mediated 

transformations that were characterized from these experiments.

4.5 Simulation Description

The modeling of the uranium bioremediation column experiments requires a specification of 

the hydrologic parameters (e.g., flow rates, porosity, dispersivity), initial solid phase 

concentrations (terminal electron acceptors, surface complexation sites, adsorbed components, 

biomass), dissolved initial and inlet concentrations (including lactate, U(VI), Fe(II), sulfate, pH, 

alkalinity, major ions), and the biogeochemical reaction network. 
4-10

The conceptual model described in Section 4.3 is translated into a quantitative specification 

for the simulation of the column experiment. For the 83-day simulation period, the two column 

experiments are operationally identical with similar behaviors (see previous effluent 

concentration plots). We chose to model Column L1 but include observations from Columns L1 

and L2 in the model comparisons. Table 4.1 summarizes the hydrologic model attributes 

including the change in flow rate at day 14. 

Table 4.1 Hydrologic parameters for Column 1: Darcy flux, pore velocity, residence time, 

porosity, and dispersivity. The flow rate to the column changed at day 14; thus, the multiple 

values represent those before and after that time point.

Parameter Value

Darcy Flux (pre/post 14 d) 0.47/0.17 cm/h

Pore Velocity (pre/post 14 d) 1.1/0.39 cm/h

Residence Time (pre/post 14 d) 9.9/27 h

Porosity 0.44

Dispersivity 2.0 cm

Table 4.2 summarizes the solid phase model specification including the initial goethite and 

phyllosilicate Fe(III) pools. As described previously, the specification of the initial concentrations 

of the goethite and phyllosilicate Fe(III) are important controls on the rise in unreacted acetate 

over the last 30 days of the column experiment. The goethite represents the crystalline Fe(III) 

whereas the phyllosilicate Fe(III) represents the structural iron in clays. The simplest 

conceptualization was to have (1) one non-limiting phyllosilicate Fe(III) pool (i.e., pool 1) 

associated exclusively with the lactate-based TEAP reaction and (2) a second phyllosilicate 

Fe(III) pool (i.e., pool 2) and goethite mineral that are bioavailable exclusively to acetate-based 

TEAP reactions. The initial mineral concentrations were calibrated to match the timing of the 

continuous and complete consumption of lactate, the initial consumption of about 70% of the 

acetate, followed by an increase in unreacted effluent acetate concentrations. The calibrated 

initial phyllosililcate Fe(III) concentrations appear to be high compared to the iron extractions 

performed on the pre- and post-biostimulation sediments. Possible explanations range from 

limitations in characterizing phyllosilicate iron to uncharacterized processes consuming lactate 

and acetate.

Surface complexation reactions for the adsorption of U(VI), Fe(II), and H+ are associated 

with a generic >FeOH surface complex. In the absence of characterization studies for these 

reactions and initial site density, we use reactions and parameters from a previous study

(Yabusaki et al., 2011). The oxidation of lactate and acetate during the biostimulation will 

produce high bicarbonate concentrations that favor the formation of aqueous uranium 

complexes over surface complexes, i.e., desorption. The expectation, based on uranium 

bioremediation field experiments in shallow aquifers is that the combination of U(VI) desorption 

and bioreduction to solid associated U(IV) will eventually deplete the adsorbed U(VI). 
4-11

Table 4.2 Initial concentrations of solid phases: goethite, initial phyllosilicate Fe(III) pools, and 

initial surface complexation site concentrations. Pools 1 and 2 are associated with the lactateand acetate-based TEAP reactions, respectively. In the model, the goethite and silicate iron in 

Pool 2 are eventually depleted resulting in the increase in effluent acetate concentration.

Solid Phase Component Concentration

Goethite 1.3E-5 mol/g

Fe(III)-phyllosilicate (Pool 1) 2.8E-3 mol/g

Fe(III)-phyllosilicate (Pool 2) 1.4E-3 mol/g

>FeOH surface complex 1.0E-5 mol/g

Biomass lactate-based FeRB 3.1E-11 mol/g*

Biomass acetate-based FeRB 3.1E-10 mol/g*

Biomass SRB 3.1E-11 mol/g*

Sorbed U(VI) 3.0E-9 mol/g

Sorbed Fe(II) 1.5E-11mol/g

Sorbed H+ 9.6E-6 mol/g

*biomass moles based on molecular formula C5H7O2N 

Initial aqueous conditions for the modeling are based on measured concentrations in the 

column experiment (Table 4.3).

Table 4.3 Initial dissolved concentrations in the column.

Component Total Concentration

pH 6.42

carbonate 12.5 mM

Ca 20.2 mM

Fe 26.4 μM

Mg 4.87 mM

Na 32.2 mM

sulfate 11.35 mM

U(VI) 13.5 μM

K 0.558 mM

The initial flow rate for column L1 was ~0.029 ml/min before being lowered to 0.011 ml/min 

at 14 days. At 0.44 porosity, the residence time increased from less than10 h to over 1 day; 

corresponding pore velocities were 1.06 cm/h and 0.39 cm/h before and after day 14 (Table 

4.1). The average influent concentrations are the same for columns 1 and 2 (Table 4.4). 
4-12

Table 4.4 Average influent concentrations from the reservoir for both column experiments.

Component Concentration

U 20.2E-6 M

Al 67.0E-6 M

Ca 20.1E-3 M

Mg 5.05 E-3 M

Fe 2.34E-6 M

K 0.785E-3 M

Na 30.5E-3 M

S 10.6E-3 M

Lactate 9.84E-3 M

pH 6.26

Alkalinity 6.28 meq/L

CO2 10.6E-6 M

4.6 Biologically-Mediated Processes: Shallow Groundwater Plume 

versus Deep ISR 

The biologically-mediated transfer of electrons to solid phase Fe(III) by FeRB and to 

aqueous sulfate by SRB provides the energy for cell maintenance, activity, and growth. 

Introducing lactate and, implicitly, acetate electron donor in solutions passing through the 

Kingsville Dome sediments stimulates FeRB that catalyze the reduction of Fe(III) mineral but 

also opportunistically catalyze the reduction of the trace amounts of aqueous U(VI) that are 

present (Lovley et al., 1991). Reduced uranium [U(IV)] becomes part of the solid phase, most 

likely as adsorbed surface coatings or uraninite forming on existing grains. 

As in the shallow aquifer sediments, aqueous Fe(II) increases very soon after biostimulation 

is initiated and is attenuated only after sulfide from the bioreduction of sulfate by SRB reacts 

with Fe(II) to from FeS mineral. This effectively removes Fe(II) from solution. Based on the 

timing of the decrease in Fe(II) at ~30 days, there is commonality with the shallow aquifer SRB 

in this lagged appearance of sulfate bioreduction, which is attributed to initially low abundance

and a slow growth rate. 

There are, however, fundamental differences between the shallow and ISL sediments. The 

first is the 10-15 day lag in U(VI) bioreduction in the Kingsville Dome column experiment. In the 

shallow aquifer sediments, U(VI) bioreduction occurs rapidly in tandem with the Fe(III) 

bioreduction. The model accounts for this lag by 1) limiting U(VI) bioreduction to acetate 

electron donor (acetate is an intermediate product of lactate oxidation) and 2) using lactate 

concentration in an inhibition term such that U(VI) bioreduction is active only when lactate 

concentrations are diminished. Another difference is the very small fraction of sulfate (10-20%) 

that is bioreduced in the Kingsville Dome sediments in the presence of large excesses of 

acetate. In the shallow aquifer sediments, the bulk of acetate consumption was attributed to 

sulfate conversion by SRB. One consequence of the largely unreacted acetate in the system is 
4-13

that there is less bicarbonate production. This may be why there is no obvious impact to the 

U(VI) concentrations due to complexation with bicarbonate that favors U(VI) desorption.

In the presence of elevated alkalinity, Fe(II), and sulfide; and lowered redox potential, calcite 

and iron sulfide secondary mineral formation are thought to be the most important volumetrically 

(Li et al., 2009). Lesser amounts of elemental sulfur, from the sulfide promoted dissolution of 

Fe(III) minerals (Fang et al., 2009), and siderite should also form.

4.7 Model Implementation 

The simulation of the 83-day lactate biostimulation for column L1 used variable time 

stepping with a maximum time step of 30 min. The 1-D model domain used to represent the 

column experiment was 10.5 cm long, comprised of 10 grid cells with uniform 1.05 cm spacing. 

The influx of lactate was based on the flow rate and influent concentrations. The reactive 

transport modeling of biostimulated reduction is an adaptation of the reaction network 

developed for the 2008 field experiment at the Rifle IFRC site (Yabusaki et al., 2011). The 

principal departure from this reaction network is the addition of the lactate-based TEAP 

reactions for Fe(III) and sulfate. 

Lactate-based phyllosilicate Fe(III) TEAP reaction

0.25CH3CH(OH)COO− + Fe(III)(ls)_pool1 = 

 0.25CH3COO- + 0.25HCO3- + 1.25H+ + 0.985Fe(II)(ls) + 

0.015Fe2+

Lactate-based sulfate TEAP reaction

0.25CH3CH(OH)COO− + 0.01562NH4

+ + 0.08595SO42- = 

 0.01562BM_SRB + 0.25CH3COO- + 0.10157H+ + 0.1719HCO3

- + 

0.08595HSIn this case, the system of equilibrium and kinetic reactions addressed 2 distinct functional 

microorganisms (i.e., FeRB and SRB), 4 TEAPs (i.e., phyllosilicate Fe(III), poorly crystalline 

Fe(III), U(VI), and sulfate) , 7 minerals (i.e., FeOOH, Fe3O4, FeS, S, calcite, siderite, uraninite), 

~100 biogeochemical species (including surface complexes). 

Key assumptions included: 1) acetate-oxidizing FeRB are solely responsible for the 

bioreduction of aqueous U(VI), 2) SRB are present and active throughout the biostimulation, 3) 

abiotic uranium reduction is negligible, and 4) geochemistry, including uranium surface 

complexation, is fully operative during biostimulation (e.g., biomass does not significantly alter 

geochemical reactivity). 

The aqueous initial and boundary conditions were lactate, U(VI), pH, bicarbonate, calcium, 

magnesium, sodium, potassium, chloride, Fe(II), and sulfate. The initial sorbed U(VI) and Fe(II) 

were assumed to be in equilibrium with their aqueous counterparts through their respective 

surface complexation models. U(VI) minerals were not included in the reaction network as they 

have not been identified in sediment analyses and are undersaturated with respect to the 

solution chemistry. Other nonaqueous initial conditions include FeRB and SRB biomass, 
4-14

goethite, 2 pools of phyllosilicate iron and surface complexation sites (Table 4.2). Both 

functional microbial groups, FeRB and SRB, are present and active during the initial period of 

biostimulation (Callister et al., 2010). The kinetics of the microbially mediated TEAP reactions is 

of the Monod type with thermodynamic control. The electron donor (i.e., lactate, acetate)

consumption rate, , is: 

( ) r

N

eA s eA eA

eA

CCs

C

m eA C

bio

C Gf CK

C

CK

C R Bs eA

∆ 















+ 















+ −= ∑ , ,

μ , 

where NeA = number of terminal electron acceptors

sC = stoichiometric coefficient of electron donor in the TEAP reaction

Cc = electron donor concentration

B = biomass concentration of microorganism mediating the TEAP reaction

CeA = terminal electron acceptor concentration

μm,eA = electron donor oxidation rate for the terminal electron acceptor

Ks,C = half-saturation coefficient for electron donor

Ks,eA = half-saturation coefficient for the terminal electron acceptor

f(ΔGr) = 1 – exp[(ΔGr- ΔGmin)/RT]

ΔGr = free energy change of the corresponding TEAP reaction

ΔGmin = minimum free energy change required to drive ATP 

synthesis

R = gas constant

T = absolute temperature

The FeRB and SRB are assumed to be attached populations whose biomass is a control on 

the rates of the TEAP reactions they catalyze. The initial FeRB and SRB biomass amounts are 

initially low (Table 4.2). As the FeRB and SRB biomass grow, the respective TEAP reaction 

rates increase commensurately. A first-order decay model (dB/dt = -kB) is used to describe the 

loss of biomass via decay. A maximum rate is used to limit the biomass-dependent reduction 

rates. 

bio RC
4-15

4.7.1 Uranium Speciation

Table 4.5 Uranium aqueous complexation reactions and formation constants.

Reaction logK (I = 0) (a)

Aqueous Species 

UO2

2+ + H2O = UO2OH+ + H+ -5.25

UO2

2+ + 2H2O = UO2(OH)2(aq) + 2H+ -12.15

UO2

2+ + 3H2O = UO2(OH)3- + 3H+ -20.25

UO2

2+ + 4H2O = UO2(OH)42- + 4H+ -32.40

2UO2

2+ + H2O = (UO2)2OH3+ + H+ -2.70

2UO2

2+ + 2H2O = (UO2)2(OH)22+ + 2H+ -5.62

3UO2

2+ + 4H2O = (UO2)3(OH)42+ + 4H+ -11.90

3UO2

2+ + 5H2O = (UO2)3(OH)5+ + 5H+ -15.55

3UO2

2+ + 7H2O = (UO2)3(OH)7- + 7H+ -32.20

4UO2

2+ + 7H2O = (UO2)4(OH)7+ + 7H+ -21.90

UO2

2+ + CO32- = UO2CO3(aq) 9.94

UO2

2+ + 2CO32- = UO2(CO3)22- 16.61

UO2

2+ + 3CO32- = UO2(CO3)34- 21.84

3UO2

2+ + 6CO32- = (UO2)3(CO3)66- 54.00

2UO2

2+ + CO32- + 3H2O = (UO2)2CO3(OH)3- + 3H+ -0.85

3UO2

2+ + CO3-- + 3H2O = (UO2)3CO3(OH)3+ + 3H+ 0.66

UO2

2+ + Cl- = UO2Cl+ 0.17

Ca2+ + UO2

2+ + 3CO32- = CaUO2(CO3)32- 25.40 (b)

2Ca2+ + UO2

2+ + 3CO32- = Ca2UO2(CO3)30

(aq) 30.55 (b)

UO2

2+ + 2Cl- = UO2Cl2(aq) -1.10

Mg2+ + UO2

2+ + 3CO32- = MgUO2(CO3)32- 26.11 (c)

UO2

2+ + SO42- = UO2(SO4)aq 3.15

UO2

2+ + 2SO42- = UO2(SO4)22- 4.14 a Values from Guillaumont et al. (2003), unless otherwise indicated.

b Bernhard et al. (2001). c Dong and Brooks (2006). 

Table 4.6 Uranium surface complexation reactions and formation constants. 

Reaction logK (estimated*)

Surface Species

SSOH + UO2

2+ = SSOUO2+ + H+ 12.28

SOH + UO2

2+ = SOUO2+ + H+ 6.95

WOH + UO2

2+ = WOUO2+ + H+ 2.74

SSOH + UO2

2+ + H2O = SSOUOOH + 2H+ 0.033

SOH + UO2

2+ + H2O = SOUOOH + 2H+ -2.12

WOH + UO2

2+ + H2O = WOUOOH + 2H+ -5.01

SSOH denoting very strong binding sites: 0.01% of total sites

SOH denoting strong binding sites: 0.1% of total sites

WOH denoting weak binding sites: 99.89% of total sites

*Values from Fang et al. (2009) 
4-16

4.7.2 Major Ion Reactions

Table 4.7 Major ion reactions and formation constants

Reaction LogK

CH3COO- + H+ = AcH 4.76

Ca2+ + H2O = CaOH+ + H+ -12.60

Ca2+ + CH3COO- = CaCH3COO+ 1.18

Ca2+ + CO3

2- + H+ = CaHCO3+ 11.33

Ca2+ + CO3

2- = CaCO3 3.15

Ca2+ + SO4

2- = CaSO4 2.31

CO3

2- + 2H+ = H2CO3 16.68

CO3

2- + H+ = HCO3- 10.33

Fe2+ + H2O = FeOH+ + H+ -9.50

CH3COO- + Fe2+ = FeCH3COO+ 1.82

Fe2+ + CO3

2- + H+ = FeHCO3+ 12.33

Fe2+ + CO3

2- = FeCO3(aq) 5.50

Fe2+ + 2CO3

2- = Fe(CO3)22- 7.10

Fe2+ + SO4

2- = FeSO4 2.25

Fe2+ + Cl- = FeCl+ 0.90

Fe2+ + 2HS- = Fe(HS)2(aq) 8.95

Fe2+ + 3HS- = Fe(HS)3

- 10.99

HS- + H+ = H2S(aq) 6.99

HS- = H+ + S2- -12.92

K+ + SO4

2- = KSO4- 0.85

Mg2+ + H2O = MgOH+ + H+ -11.79

CH3COO- + Mg2+ = MgCH3COO+ 1.14

CO3

2- + Mg2+ + H+ = MgHCO3+ 11.40

CO3

2- + Mg2+ = MgCO3(aq) 2.98

Mg2+ + SO4

2- = MgSO4(aq) 2.25

Na+ + CH3COO- = NaCH3COO -0.18

Na+ + CO3

2- + H+ = NaHCO3(aq) 10.08

CO3

2- + Na+ = NaCO3- 1.27

Na+ + SO4

2- = NaSO4- 0.70

NH4

+ = NH3(aq) + H+ -9.25

NH4

+ + SO42- = NH4SO4- 1.11

SO4

2- + H+ = HSO4- 1.99

H2O = OH- + H+ -14.00

>FeOH + H+ = >FeOH2

+ 7.47

>FeOH = >FeO- + H+ -9.51

>FeOH + Fe++ = >FeOFe+ + H+ -5.00

>FeOH + Fe++ + H2O = >FeOFeOH + 2H+ -11.96

CO2(aq) = CO2(g) 2.1
4-17

4.7.4 Kinetic Reactions

Table 4.8 Kinetic reactions, rate laws, and reaction parameters. 

Kinetic Reactions Rate Laws* Parameters

Ca2+ + HCO3

-

= Calcite(s) + H+

( )

 ( ) 







<Ω−Ω

 

−Ω ≥Ω

=

− 1 for 1

1 for 1

ii i i

i i i

i

Qk

k

r

ki = 1.644×10-7 Md-1

k-i =1.37×10-6 d-1

Ksp = 71.48

Fe2+ + HCO3

- = Siderite(s) + H+

( )

 ( ) 







<Ω−Ω

 

−Ω ≥Ω

=

− 1 for 1

1 for 1

ii i i

i i i

i

Qk

k

r b 

ki = 1.37×10-6 Md-1

k-i = 1.37×10-7 d-1 

Ksp = 0.6504

Fe2+ + HS-= FeS(s) + H+

( )

 ( ) 







<Ω−Ω

 

−Ω ≥Ω

=

− 1 for 1

1 for 1

ii i i

i i i

i

Qk

k

r

ki = 5.11×10-6 Md-1

k-i = 6.85×10-6 d-1 

Ksp = 2.25e-4 

0.125 CH3COO-+ 

Fe(III)(ls)_pool2 + 0.5 H2O = 

0.25 HCO3

- + 1.0 Fe(II)(ls) + 

1.125H+

r [acetate][lactate] = μi μi = 1.6×104 M-1

d-1

0.125 CH3COO− + FeOOH(s) + 

1.875 H+ = Fe2+ + 1.5 H2O + 

0.25 HCO3

−

][ [ ][ ]

K lactate

K r FeOOH Ac

i

i i + = μ

μi = 40.0 M-1d-1

Ki = 1.3×10-3 M 

0.125CH3COO- + 0.3538H2O + 

0.0113NH4

+ + 0.3875UO22+ = 

0.0113BM_iron + 0.855H+ + 

0.1938HCO3

- + 0.3875UO2(s)

















+ 















+ 















+ = ][ , ,

, K lactate

K

CK

C

CK

C Br

i

i

s eAeA

eA

cCs

c μ m eA

μm,eA = 0.8 Md-1

Ks,C = 1×10-9 M 

Ks,eA = 5.0×10-5 M 

Ki = 3.0×10-5 M 

0.25CH3CH(OH)COO− + 

0.01562NH4+ + 0.08595SO42- = 

0.01562BM_lac + 

0.25CH3COO- + 0.10157H+ + 

0.1719HCO3

-

+ 0.08595HS-

















+ 















+ =

s eAeA

eA

cCs

c

m eA CK

C

CK

C Br

, ,

μ ,

μm,eA = 2.0×10-2 Md-1

Ks,C = 5.0×10-4 M 

Ks,eA = 1.0×10-5 M 

0.25CH3CH(OH)COO− + 

Fe(III)(ls)_pool1 = 

0.25CH3COO- + 0.25HCO3- + 

1.25H+ + 0.985Fe(II)(ls) + 

0.015Fe2+

r Fe _[ lactatels ][][ BM lac]_ = μi μi = 2.0×102 M-2

d-1

*

Reaction-rate formulations for calcite, siderite, and iron sulfide are functions of the saturation state follow Hunter et al. (1998). 

Nomenclature

i = the i-th mineral phase

i

r = the rate of the mineral reaction

i k = the mineral precipitation rate [ML-3

T-1]

i k− = the mineral dissolution rate [ML-3

T-1]

Qi = the concentration of the i-th mineral phase [ML-3

]

Ωi = the saturation index of the i-th mineral phase

Ksp = saturation constant
4-18

r = the rate of bioremediated reaction

Cc = electron donor concentration

CeA = terminal electron acceptor concentration

μm,eA = electron donor oxidation rate for the terminal electron acceptor

Ks,C = half-saturation coefficient for electron donor

Ks,eA = half-saturation coefficient for the terminal electron acceptor

μi = rate coefficient

ki = Inhibition constant [M]

[Ac] = Acetate concentration [M]

[FeOOH] = Goethite concentration [M]

[Fe_ls] = Layer silicate iron (III) concentration [M]

[lactate] = Lactate concentration [M]

[BM_lac] = Lactate related biomass concentration [M]

4.8 Biostimulation Modeling Approach for the Column Experiments

The goal of the modeling is to interpret the observed behavior of the Kingsville Dome 

sediment column experiments by developing a reaction network and parameterization to 

simulate the bioreduction of sulfate, Fe(III), and U(VI) in the lactate biostimulation column 

experiments using Kingsville Dome sediments. While column experiments typically use a finergrained subset of the field particle size distribution, the sediments recovered from the Kingsville 

Dome drilling were sand size and smaller (see Table 3.6), which allowed direct use in the 

column studies. 

In the absence of a well-instrumented biorestoration field experiments at an ISR site, the 

column experiments were intended to provide basic information on the potential for 

biorestoration of uranium ISR sites. The recognition here and stated in Chapter 2 is that 

laboratory conditions are not necessarily representative of the field conditions that control the

behavior of uranium and other components of interest at ISR sites. In this case, preserved postISR sediments recovered from the Kingsville aquifer are used in column experiments where 

indigenous microorganisms are stimulated with lactate to catalyze the conversion of aqueous 

U(VI) to immobile U(IV). While lactate is commonly used as an electron donor, it should be 

recognized that the specific electron donor – terminal electron acceptor – microorganism – 

geochemistry combination yields specific behaviors. The first degradation product of lactate is 

acetate, which is also a commonly used electron donor. So in general, there will need to be two 

sets of microbially-mediated TEAP reactions, one for lactate and one for acetate (Table 4.8). 

4.9 Column Simulation Results

In the model, lactate is consumed in the sulfate and phyllosilicate Fe(III) TEAP reactions. 

The continuously influent 10 mM lactate is rapidly consumed over the first 14 days at an 
4-19

increasing rate (Figure 4.5). The rapid drop in effluent lactate concentrations between 12 and 

14 days is due to a decrease in the influent flow rate from 0.03 ml/min to 0.011 ml/min. The 

increase in residence time from 10 to 27 h allowed nearly complete lactate consumption by day 

14. From day 14 to the end of the 83-day biostimulation stage of the experiment, the 

continuously influent 10 mM lactate is essentially completely consumed during the 1.1 day

transit through the column.

Acetate is being progressively produced within the column as lactate is transported by the 

flow and oxidized by FeRB and SRB via the lactate-based phyllosilicate Fe(III) and sulfate 

TEAP reactions, respectively. From day 15 to 55, effluent acetate concentrations are ~3 mM. 

This implies that ~7 mM acetate, of the 10 mM acetate produced by the complete oxidation of 

the influent 10 mM lactate, was being consumed. In the model, there are three acetate-based 

TEAP reactions. Two are Fe(III) reactions: goethite (representative of the bioavailable 

crystalline Fe(III) terminal electron acceptor) and phyllosilicate Fe(III) (iron containing clays). 

The third is for the U(VI) TEAP reaction. All three modeled TEAP reactions are mediated by 

FeRB. After day 55, acetate concentrations rapidly rise to 9 to 10 mM in column L1 and 7-8 mM 

increasing to over 9 mM in column L2 (Figure 4.6). The model accomplishes this behavior 

through the sequential depletion of the bioavailable phyllosilicate Fe(III) and goethite. In this 

case, the acetate-based phyllosilicate Fe(III) TEAP reaction rate is faster resulting in 

progressive depletion of bioavailable Fe(III) from the front end of the column as compared to the 

acetate-based goethite TEAP reaction which is slower resulting in higher depletion at the 

effluent end of the column. 

Figure 4.5 Lactate concentrations in column effluent: comparison of model and observations from 

columns 1 and 2.
4-20

Figure 4.6 Acetate concentrations in column effluent: comparison of model and observations 

from columns 1 and 2.

There is an initial transient pulse of effluent ~60 μM Fe(II) at 2 days that is not captured by 

the model (Figure 4.7). This can occur when sediments are first exposed to flowing solution

and is observed in column L5 (Figure 3.10) which did not undergo bioreduction. The column 

experiment observations appear to settle down after 4 days whereupon the effluent Fe(II) 

steadily increases with the rate of increase slowed by the drop in flow rate from 0.03 ml/min to 

0.011 ml/min at day 14. Peak Fe(II) is ~300 μM occurring at 30 days for column L1 and 25 days 

for column L2. Prior to the drop in flow rate, the modeled effluent Fe(II) increases relatively 

slowly to 34 μM. After the drop in flow rate the additional residence time allows a higher rate of 

Fe(III) bioreduction resulting in a relatively rapid climb in predicted effluent Fe(II) to ~300 μM 

after 22 days. The decline in the Fe(II) after the peak is attained occurs during a period of 

steady consumption of 10 mM lactate and 7 mM acetate (15 to 55 days). During this period, the 

principal pools of bioavailable terminal electron acceptors (i.e., goethite, phyllosilicate Fe(III) and 

sulfate) in the model are not limiting. The progressive increase in SRB biomass has resulted in 

an increasing amount of sulfide being produced through the sulfate TEAP reaction to react with 

the Fe(II) to form FeS mineral. This is because the sulfate TEAP reaction rate law has a 

biomass dependent term. As more and more sulfide is produced, more and more Fe(II) is 

removed from solution until day 45, when the Fe(II) is largely removed from solution. While the 

model does reproduce the general trend of the observed Fe(II) breakthrough, including the peak 

effluent Fe(II) concentration, the temporal width of the simulated pulse is not as wide as 

observed in columns L1 and L2. In the model, there is a trade-off between capturing the peak, 

which is the net result of the three Fe(III) (goethite and 2 phyllosilicate pools) TEAP reactions 

and rates and one sulfate TEAP reaction and rate; and capturing the total amount of Fe(II) 

liberated by bioreduction. This is further constrained by the eventual depletion of the 

bioavailable acetate-based goethite and Fe(III) terminal electron acceptor pools. 

In the model there is significantly more Fe(II) produced than is observed in solution. The 

Fe(III) TEAP reactions reduce 4 moles and 8 moles of Fe(III) for every mole of lactate and 

acetate consumed, respectively. That means that the equivalent of at least ~75 mM Fe(II) is 

being produced at the height of lactate and acetate consumption. The phyllosilicate Fe(III) is 

generally considered to be reduced in place in the layer silicate clay structure. Conversely, the 
4-21

goethite in the acetate-based Fe(III) TEAP reaction is assumed to be reductively dissolved, 

liberating Fe(II) in solution. Solid phase reactions involving the formation of secondary minerals 

(e.g., FeS, siderite) and Fe(II) surface complexation effectively remove the bulk of Fe(II) 

liberated into solution. 

Figure 4.7 Fe(II) concentrations in column effluent: comparison of model and observations from 

columns 1 and 2.

In the model, the bioreduction of sulfate is lactate-based and active from the beginning of 

the biostimulation. Prior to the drop in the flow rate, the modeled sulfate bioreduction removed 

a nearly constant 0.75 mM sulfate from the influent solution. The modeled sulfate removal 

increased to ~1.5 mM after the drop in flow rate, which allowed more reaction time. However, 

the modeled sulfate removal slowly decreased to ~1.0 mM by the end of the experiment, due to 

slowly increasing lactate consumption in the Fe(III) TEAP which made less lactate available for 

the sulfate TEAP reaction. Over the course of the experiment, the removal of sulfate is always 

less than 2 mM of the 10.65 mM influent sulfate concentration (Figure 4.8). As the bioreduction 

of one mole of sulfate results in one mole of sulfide, even 1-2 mM of sulfide is sufficient to react 

with essentially all of the aqueous Fe(II) in the column. The stoichiometry of the sulfate TEAP 

reaction consumes nearly 3 moles of lactate for every mole of sulfate reduced. That means that 

3 to 6 mM of the 10 mM influent lactate is consumed in the sulfate TEAP. The remainder of the 

lactate is consumed in the phyllosilicate Fe(III) TEAP. 
4-22

Figure 4.8 Sulfate concentrations in column effluent: comparison of model and observations 

from columns 1 and 2.

The continuous influent 20 μM U(VI) is attenuated by precipitation and adsorption. In the 

model, U(VI) bioreduction via the acetate-based TEAP reaction is the dominant removal 

mechanism. Until day ~14, effluent U(VI) exhibits little attenuation from the influent 20 μM 

concentration. The onset of U(VI) attenuation corresponds to the time when lactate becomes 

largely depleted (Figure 4.9). The model accomplishes this lag in U(VI) removal by using the 

presence of lactate to inhibit the U(VI) bioreduction. Thus, U(VI) bioreduction will be significant 

only where lactate has been sufficiently oxidized. The rate of U(VI) bioreduction increases with 

time until ~day 40, when U(VI) in the effluent becomes negligible. In the model, the increasing 

rate of U(VI) bioreduction is the result of a biomass-dependent rate law for the acetate-based 

U(VI) TEAP reaction. FeRB biomass increases with the continuously influent lactate. In the 

model, uraninite is produced throughout the column and can account for up to 99.98% of the 

solid phase uranium. This is because of 1) the reduction kinetics, which allow more reaction in 

a transported fluid parcel with longer residence time (equivalently deeper transport into the 

column), and 2) the removal of U(VI) from solution via bioreduction lowers aqueous 

concentrations, which drives more U(VI) desorption. 

 
4-23

Figure 4.9 U(VI) concentrations in column effluent: comparison of model and observations 

from columns 1 and 2.

Model-predicted spatial distributions of solid-associated components at the end of the 83-

day biostimulation period are shown in Figures 4.10-4.13. There is considerable spatial 

variation in these concentrations that reflects the microbial reaction kinetics. The introduction of 

lactate allows the lactate-based bioreduction reactions to consume phyllosilicate Fe(III) and 

sulfate, with the highest rates nearer the influent end of the column where the lactate 

concentrations are the highest. Consequently, the lactate-based phyllosilicate Fe(III) is 

progressively consumed from the influent end of the column. As lactate is transported through 

the column these biologically-mediated lactate oxidation reactions produce acetate. As a result, 

the acetate-based bioreduction of goethite and U(VI) occurs deeper into the column away from 

the column inlet. In this case, goethite is progressively depleted beginning at the column outlet.

At the end of the 83-day simulation, FeS mineral has precipitated throughout the column 

with the highest abundance nearer the column inlet (Figure 4.10). At this location, the lactatebased sulfate bioreduction has produced high concentrations of sulfide which react with the 

aqueous biogenic Fe(II) to form the FeS mineral. This result is qualitatively consistent with the 

Table 3.8B trend of higher solid phase Fe(II) and Table 3.8C trend of higher solid phase S 

concentrations near the influent end of the column.
4-24

Figure 4.10 FeS(s) abundance as a function of distance from column entrance at the end of the 

83-day biostimulation. (1 dm = 10 cm)

During the 83-day simulation, adsorbed U(VI) desorbs from sediments throughout the 

column but is preferentially desorbed beginning from the effluent end of the column (Figure 

4.11). This is due to 1) the progressive increase in bicarbonate concentrations as lactate and 

acetate continuously oxidize while being transported through the column, and 2) the acetatebased bioreduction of aqueous U(VI). In the first case, the elevated bicarbonate concentrations 

thermodynamically favor the formation of U(VI) aqueous complexes over surface complexes. In 

the second case, the bioreduction of aqueous U(VI) to solid-associated U(IV) effectively 

removes U(VI) from solution, which also drives further desorption of adsorbed U(VI). Note that 

towards the end of the experiment when goethite and phyllosilicate Fe(III) pools for acetate 

bioreduction are being depleted, acetate oxidation is limited, which decreases the bicarbonate 

production and U(IV) conversion.
4-25

Figure 4.11 Sorbed U(VI) abundance as a function of distance from column entrance at the end 

of the 83-day biostimulation.

At the end of the 83-day simulation, solid-associated U(IV) has a peak concentration 2 cm 

from the effluent end of the column (Figure 4.12). Since effluent acetate concentrations at this 

time are approaching the stoichiometric conversion from lactate (i.e., acetate is largely 

unreacted), the U(IV) peak away from the column outlet is the result of limited availability of 

aqueous U(VI) beyond this location. This is consistent with the depletion of sorbed U(VI) 

towards the end of the column (Figure 4.11). The solid phase analyses of the postbiostimulation column sediments in Table 3.8A identified the peak U(IV) concentrations in the 

upgradient half of the column. This can happen if the U(VI) bioreduction rate is higher than in 

the model.
4-26

Figure 4.12 U(IV) concentration as a function of distance from column entrance at the end of 

the 83-day biostimulation.

The bioreduction of phyllosilicate Fe(III) leads to biogenic Fe(II) that largely remains in the 

layer silicate structure. Figure 4.13 shows the near linear decrease in phyllosilicate Fe(II) from 

the beginning to the end of the column at the end of the 83-day simulation. This generally 

reflects the lactate-based phyllosilicate Fe(III) TEAP reaction rate which is fastest at the column 

inlet where the highest lactate concentrations are found and becomes progressively lower as 

lactate is consumed as it is transported through the column. The Fe(II) in Table 3.8B was also 

higher at the inlet but generally flat after that. Concentration magnitude based on the 

extractions was generally lower in the post-biostimulation solid phase extractions than in the 

model simulations. Possible explanations range from difficulty extracting phyllosilicate iron to 

the consumption of electron donor by processes other than the modeled TEAPs. 
4-27

Figure 4.13 Fe(II) silicate abundance as a function of distance from column entrance at the end 

of the 83-day biostimulation.

4.10 Biostimulation Studies: Deep ISR vs Shallow Aquifer 

The most important findings from the studies of the biostimulation column experiments using 

the preserved Kingsville Dome post-ISR sediments were 

1. The microbial community in the post-ISR sediment column could be stimulated with 

lactate resulting in essentially complete removal of continuously influent 20 μM U(VI) 

over the ~1-day column transit time. The viability of the anaerobe populations 

putatively responsible for the bioreduction of U(VI) to immobile, sediment associated 

U(IV) is significant in light of the prolonged exposure to oxygen and other 

components of the lixiviant during ISR operation. Oxygen inhibits the activity of 

these organisms and can, in some cases, be toxic. Thus, the ability to respond 

within the experimental time scales to biostimulation after the disruption imparted by 

ISR extraction is notable.

2. The response of the microbial community in the Kingsville Dome sediments to lactate 

biostimulation is distinctly different from that observed in shallow aquifer sediments. 

In the biogeochemical modeling, these differences were addressed with inhibition 

terms, biomass-dependent reaction rates, and electron donor-specific (a) TEAP 

reactions and (b) bioavailable terminal electron acceptor pools. 
4-28

Acetate. The first oxidation product in the lactate-based TEAP reactions is acetate, which is 

another electron donor that can also drive the microbially-mediated reactions. After 14 days of 

continuous biostimulation, the complete continuous consumption of the 10 mM influent lactate 

electron donor produces 10 mM of acetate according to the stoichiometry in the modeled TEAP 

reactions. Initially, 70% of the acetate is consumed in TEAP reactions but by the end of the 

experiment, nearly all the acetate leaves the column unreacted. This is a complete departure 

from the shallow alluvial continuous acetate biostimulation experiments where acetate initially 

drives Fe(III) and U(VI) bioreduction but then is nearly completely consumed mole for mole by 

sulfate bioreduction. The diminishing ability of the microbial community in the Kingsville Dome 

sediments to utilize acetate was modeled as the progressive depletion of Fe(III) pools of limited 

size that are bioavailable only to acetate-based TEAP reactions. 

Sulfate. When sulfate reduction is the dominant TEAP in the Rifle shallow aquifer system, the 

acetate electron donor will continue to be consumed as long as sulfate is not limiting. While the 

ambient sulfate concentrations ( ~10 mM) in the ISR sediment column experiments are 

comparable to the Rifle aquifer, only a small amount (1-2 mM) of sulfate is removed from 

solution via bioreduction despite the presence of excess acetate electron donor. In the model, 

this insensitivity of sulfate reduction to acetate is addressed by limiting the TEAP reaction 

mediated by sulfate reducing bacteria to lactate electron donor. Since much less sulfate is 

reduced to sulfide, less electron donor is consumed, and less biomass is produced. This could 

be viewed as a benefit, in that more of the electron donor is being used by the microorganisms 

responsible for catalyzing the reduction of U(VI). Conversely, there is less sulfide produced by 

sulfate reducing bacteria, which decreases the solid phase reservoir of lowered redox potential 

that contributes to the restoration and maintenance of baseline conditions. 

Fe(III) Minerals. The bioreduction of iron (oxy)hydroxide (i.e., goethite) and phyllosilicate Fe(III) 

(iron-bearing clays) account for the bulk of the electron donor consumed in the modeled 

experiments. The rise and fall of Fe(II) in the first 40 days of column effluent is similar to the 

biostimulation of iron-reducing bacteria in shallow alluvial aquifer systems where the initial 

increase in aqueous Fe(II) is eventually attenuated by the formation of FeS from reaction with 

increasing sulfide concentrations. In the model, Fe(II) is liberated by the reductive dissolution of 

the crystalline Fe(III), whereas the Fe(III) in the clays is reduced in place with minimal release of 

Fe(II) into solution. In this case, the iron-bearing clays are the dominant Fe(III) TEAP. Solid 

phase iron characterization indicated less electron donor consumption via TEAP reactions and 

more consumption by undetermined non-TEAP processes than simulated. While this does not 

alter the principal conceptual processes simulated, it does underscore a need to further explore 

alternative conceptual models for the column behaviors. 

U(VI). Unlike the shallow alluvial aquifers in which the bioreduction of U(VI) (and Fe(III)) 

responds almost immediately to the biostimulation, there is a 14-day lag in the Kingsville Dome 

sediment columns that is coincident with complete lactate consumption. In the model, the 

presence of lactate inhibits the U(VI) bioreduction which is attributed to acetate. Thus, U(VI) 

bioreduction is 1) acetate-based, despite acetate being largely unreacted by day 60, and 2) 

inhibited by the presence of lactate. U(VI) removal from solution progressively increased after 

the 14-day lag period until negligible amounts of U(VI) were present in the column effluent after 

40 days of lactate biostimulation.
5-1 

5 Field-Scale Flow and Transport Modeling 

5.1 Overview

The conceptualization of processes identified for the laboratory columns using Kingsville 

Dome sediments was extended to the field scale using a hypothetical confined aquifer. The 

general attributes of the site were selected from a range of values based on ISR sites and are 

listed in Table 5.1. The scenario being modeled is that this section of the confined aquifer was 

previously the subject of uranium ISR. In this scenario, the elevated uranium concentrations 

and major ion chemistry (see Tables 4.2 and 4.3) are based on field observations of 

groundwater during the drilling of the post-ISR sediments used in the column experiments. 

Table 5.1 Hypothetical sand-sandstone aquifer attributes. 

Parameter Value

Thickness 10 m

Hydraulic conductivity 1.4 m/d

Specific storage 1.0E-5 1/m

Porosity 0.15

Depth below water table 104 m

Water pressure 10 atm/103 mH2O

Restoration is based on the delivery of 10 uM lactate to the 10 m thick confined aquifer 

interval using multiple injection wells coordinated with a single central production well. The 

most common injection/pumping patterns are five- and seven-spot (Lusher, 2003), although ore 

body geometry and/or surface topography may give rise to other patterns such as direct line. 

For this hypothetical scenario, we chose a five-spot well pattern with 4 injection wells on the 

corners of a 30.5 m square with a central recovery well in the middle of the square. Distance 

between the center extraction well and corner injection wells was 21.6 m. This 5-spot scenario 

is a highly simplified and idealized problem geometry designed to focus on the basic interplay 

between flow and reaction processes in a field setting. In a typical ISR well field, the well 

network will be much larger and more complex, with distances between wells varying as a 

function of the size of the mineralized zone, permeability, pumping rates and drawdown. 

Spacing between injection wells is generally 40 to 150 feet. Our single 5-spot analysis can be 

considered a “building block” element of a larger network of 5-spot patterns (see Figure 1.3) 

that are typically used to cover the zone of economic uranium extraction. Table 5.2 contains 

the modeled 5-spot well specifications, including pumping rates and drawdown. 
5-2 

Table 5.2 Well attributes for 5-spot pattern. 

Parameter Value

Well diameter 6 inches

Injection wells 4

Production wells 1

Injection rate 10 gpm

Production rate 40.4 gpm

Drawdown 54 feet

The flow modeling used a two-dimensional, depth-averaged model domain to analyze the 

drawdown, time to near steady-state, capture zone, and sensitivity to gradient and dispersion. 

While the idealized conditions for the modeling do not include the impact of spatially variable

material properties, some insights on potential impacts of heterogeneities were identified 

through the dispersivity tests. 

The conceptualization of the ISR-mined region is based on an idealized spatial distribution 

of chemical species (including uranium, iron, sulfur), biological species (including initial FeRB 

and SRB biomass, and mineralogy (including the crystalline and phyllosilicate iron terminal 

electron acceptors, as well as the >FeOH surface complexation sites). In this case, the 2-D 

depth-averaged domain is for a confined aquifer with Dirichlet hydraulic head boundary 

conditions. These boundary conditions drive a linear uniform regional gradient. The 5-spot 

pattern of injection wells is a set of hydrologic sources and sinks that perturb the plane of the 

regional piezometric surface. The porosity, hydraulic conductivity, and dispersivity are uniformly 

assigned to the entire model domain. As in the modeling of the column experiment, the initial 

aqueous solute speciation and chemical reactions are used to assign the adsorbed U(VI), Fe(II), 

and H+ concentrations. The system is assumed to be initially anaerobic. While the use of 

oxygen in the lixiviant may have left residual amounts in the aquifer, it takes less than 1 mM of 

electron donor to remove all the accessible oxygen (Yabusaki et al., 2010). The assumption 

here is that biostimulated removal of oxygen is sufficiently efficient to deal with parts of the 

aquifer that are less accessible to the primary flow paths. The use of pumping and injection 

wells complicates the concept of a pore volume that works so well with column experiments. 

While the zone of influence for the 5-spot pattern certainly has a pore volume associated with it, 

the amount of water transported through any part of the zone is location-dependent. Thus, we 

use the time it takes for the longest travel time to reach the pumping well as a measure of cycle 

time. The amount of groundwater that must be pumped is a significant component of the overall 

cost of decommissioning a site. 

5.2 Field-Scale Coupled Process Simulator 

eSTOMP, the scalable parallel processing version of the STOMP subsurface simulator 

(White and Oostrom, 2006), was used to model the coupled saturated flow and transport 

processes and the large number of reactive components for the two-dimensional field-scale 
5-3 

simulations of uranium ISR site restoration. eSTOMP was selected for the field scale coupled 

process simulations because the high-resolution reaction dominated computations could be 

executed on massively parallel computers. In this case, the operator splitting reaction solver is 

identical to the one employed in HYDROGEOCHEM. The eSTOMP simulator solves transient 

flow and multicomponent biogeochemical reactive transport problems in the subsurface 

environment in any dimensionality. The governing flow and transport equations are the partial 

differential equations (PDEs) for the conservation of water and solute mass. All boundary 

conditions, sources, and sinks can be time variant and applied selectively over the boundary 

surfaces. The water mass conservation equation, shown below, equates the time rate of 

change of water mass within a control volume with the flux of water mass crossing the control 

volume surface. Flow of fluid phases is computed from Darcy’s law. 

The solute conservation equation, shown below, equates the time rate of change of solute 

within a control volume with the flux of solute crossing the control volume surface. In the 

eSTOMP simulator, solute is partitioned among the fluid and solid phases assuming 

thermodynamic and geochemical equilibrium conditions. Solute transport occurs by advection 

and diffusion-dispersion through the aqueous phase. 

Equilibrium reactions (i.e., mass action equations relating species activities through an 

equilibrium constant) are represented as
5-4 

Kinetic equations define kinetic components where a stochiometrically weighted sum of 

species concentrations vary in time according to a weighted sum of kinetic rates: 

A variety of popular reaction rate models are available including those based on simple first 

order reactions, transition state theory, and Monod kinetics.

eSTOMP uses one-sided communication and a global shared memory programming 

paradigm from the Global Array Toolkit (GA) library (Nieplocha et al., 2006) for scalability, 

performance, and extensibility on massively parallel processing computers. The approach is 

compatible with the more commonly used Message Passing Interface (MPI) (Message-PassingInterface-Forum, 2009) that is used by the PETSc (Balay et al., 2010) global implicit solver in 

eSTOMP. The principal benefit to this parallel processing approach is that simulations with 

highly resolved process and property detail can be performed on a massively parallel computer 

in hours instead of weeks on a desktop computer. These efficiencies enabled rapid turnaround 

of the multiple test scenarios that were simulated.

Written in Fortran-90, eSTOMP solves the conservation equations for compressible flow and 

multicomponent reactive transport based on finite-volume discretization. A Newton-Krylov 

solution procedure is used for the nonlinear global implicit formulation of the governing flow and 

transport PDEs. Solute mass conservation and reactions are sequentially solved following the

solution of the flow equations. Kinetic and equilibrium reactions are represented as a system of 

ordinary differential and algebraic equations (DAEs). The nonlinear system of DAEs is 

formulated with the backward Euler method and Gauss-Jordan matrix decomposition. NewtonRaphson iteration is used to solve the reaction system of equations. 

5.2.1 Physical System

Attributes of the model used in this study were chosen to be representative of typical ISR 

sites with sandstone formations containing uranium roll front deposits. The single, 10 m thick 

layer aquifer was confined by no-flow lower and upper boundaries to represent the bottom- and 

top-confining geology, respectively. The 2-D plan view model constrains all flow to the

horizontal. Hydraulic conductivity was specified at 1.4 m/d, porosity at 0.15, and the specific 

storativity (meters of water released per m decrease in head) was 1.0E-5 1/m. Background 

pressure was set at 10 atm, or about 340 feet of hydraulic head. A typical pumping rate of 10 

gpm was used for each injection well, and 40.4 gpm (4 injection wells x 10 gpm +1%) was used 

for the center extraction well. A conservative (non-reacting, unretarded) tracer with a nominal 

concentration of 1.0 was included in the injectate of the four injection wells. 
5-5 

A sensitivity analysis was conducted on this baseline specification to evaluate the 

significance of model grid resolution and extent, regional hydraulic gradient magnitude and

direction, and dispersion. Tables 5.3 and 5.4 list these inputs for the sensitivity analysis. The 

east-to-west and southwest-to-northeast flow directions represent the possible extremes of 

background flow in relation to this well field. Because of well field symmetry, no other flow 

directions are needed.

Table 5.3 Model domain and sensitivity parameters. 

Parameter Test A Test B

Regional hydraulic gradient magnitude 5e-3 5e-4

Regional hydraulic gradient direction West to east Southwest to northeast

Diffusivity and dispersivity Diff = Disp = 0 Diff=water, Disp = 10 m

Table 5.4 Grid testing attributes for Grid 1 and 2.

Attribute Grid 1 Grid 2

Extent 195.75 m x 195.75 m 395.75 m x 395.75 m

Dimensions in Grid Cells 161 x 161 277 x 277

Total Cells 25921 76729

5.2.2 Model Flow System

The five-spot well pattern forms a square, and the entire model domain was also a square, 

with the objective of having enough buffer area around the well field such that boundary 

conditions would not unduly influence the water flow and transport in the vicinity of the wells. 

Cell size was variable to reduce the number of computational nodes in the model and save 

runtime. Model cells (each centered on a computational node) ranged in size from 0.15 m (~6 

in) to 40 m on a side. The minimum cell size of 0.15 m is similar to the diameter of typical ISR 

production wells in the field. Cell size was set to increase with distance from the wells so that 

the grid was fine in the dynamic and high gradient regions of the flow system, and relatively 

coarse in the other regions of the flow system. 

Grid Testing. Two grids were used to assess 1) the adequacy of the domain size (i.e., 

minimize boundary effects on the flow field around the five-spot well pattern), 2) grid resolution

(i.e., capture spatially variable flow and transport features), and 3) convergence of the simulated 

field variables. The larger grid used more nodes (smaller growth in spacing) in the vicinity of the 

wells, as well as having more cells to make a wider buffer away from the well field. The smaller 

grid was 195.75 m on a side, and had 161 nodes in the x- and y-directions, for a total of 25,921 

nodes (Figures 5.1 and 5.2). The larger grid was 395.75 m on a side, and had 277 nodes in 
5-6 

each direction, for a total of 76,729 nodes (Figures 5.3 and 5.4). In both grids the well field was 

placed at the center of the model. Grid 1 (smaller) was the primary grid used in production runs. 

Grid 2 served as a check on the adequacy of grid 1; if the larger extent and greater number of 

nodes in grid 2 do not yield significantly different results from grid 1, then grid 1 should be 

adequate. 

Figure 5.1 Model grid 1. The square domain is 196 m on a side, has the same spacing in the xand y-directions, and contains 25,921 nodes. Blue dots are injection wells, red dot is extraction 

well.
5-7 

Figure 5.2 Model grid 1 center area, showing small cells used in the vicinity of the wells.
5-8 

Figure 5.3 Model grid 2. The square domain is 396 m on a side, has the same spacing in the xand y-directions, and contains 76,729 nodes. Blue dots are injection wells, red dot is extraction 

well.
5-9 

Figure 5.4 Model grid 2 center area, showing small cells used in the vicinity of the wells.

Boundary conditions were constant head type for groundwater (i.e., Dirichlet), and advection 

outflow type for the tracer. These respective conditions defined a steady and uniform 

background hydraulic gradient and allowed the tracer to leave the model in the pumped 

groundwater. In the two-dimensional, depth-averaged model, the injection wells were 

designated point sources of injectate with tracer concentration of 1.0, and the extraction well 

was a designated sink where groundwater and tracer were removed. Each source or sink was 

represented by a single node in the model, having the smallest cell size, 0.15 m x 0.15 m. Two 

versions of "spinup" or attainment of a steady state in water flow prior to tracer injection, were 
5-10

tried. The first one involved running the model to 20 years with just the background, regional 

flow. The second one included well pumping (water only) in addition to the background regional 

flow. In both versions, most of the change in the flow field occurred in the first two days, and all 

change in groundwater velocities was zero within 200 days. Since the time period of interest for 

evaluating tracer transport was a year or more, either version was sufficient for the modeling 

objectives. 

5.2.3 Flow and Transport Results

The operation of the five-spot wells results in local perturbations from the regional gradient 

of the piezometric head surface. For both grids, these perturbations are shown in Figure 5.5 for 

the 4 combinations of the1) 0.005 and 0.0005 regional hydraulic gradients and 2) gradient 

direction parallel and diagonal to the five-spot square. Within a given grid, the head change is 

identical, indicating proper model execution, i.e., consistency with the principle of superposition 

in a linear PDE. Away from the well field, there will be head differences between the two grids 

because of the different distances between the five-spot wells and the model boundaries. 

These differences, however, are very small and the spatial patterns of head change are 

essentially the same in both grids.
5-11

Change in Hydraulic Head (m)

Figure 5.5 Top: change in hydraulic head in meters caused by pumping wells, in comparison to 

the background flow state, grid 1. Bottom: change in hydraulic head caused by pumping wells, 

in comparison to the background flow state, grid 2.

As a first check on the transport away from the injection wells, the paths of released 

particles were computed and plotted in Figure 5.6. Routing was based on advection by the 

steady-state velocity field with the five-spot well pattern in operation. Sixty particles were placed 

in a circle around each injection well, spaced 6 degrees apart, at a radial distance of 0.32 

meters from the well node. From these initial positions, the particles were moved in small 

increments according to the groundwater velocity field. Particles were either captured by the 

central extraction well, transported out of the model domain at the boundaries, or in a few 

special cases, stagnated in zones where the regional flow and 5-spot operations canceled. 

Figure 5.6 shows that all particles were captured by the extraction well in cases with the smaller 

(5e-4) background gradient. However, the larger background gradient resulted in some 

particles escaping the well field and moving to the model boundaries. These particle excursions 

occur where initial movement caused by injection puts the particles out in areas where the 

background flow has more influence than the extraction. Such particles are initially pushed in a 

direction away from the extraction well, and in a direction close to that of the background flow. 

These results were the same for both grids. It should be emphasized that these flowlines 
5-12

represent the effects of advection only--that is, movement according to the bulk groundwater 

flow, and with no dispersion.

Figure 5.6 Top: paths of particles initially placed around injection wells, and color contours of 

head change, grid 1. Bottom: paths of particles initially placed around injection wells, and color 

contours of head change, grid 2.

The distribution of travel times for the 60 particles initially placed around each injection well

is large because of the different particle paths with spatially variable velocities. In this model, 

hydrologic properties are homogeneous; thus, the spatially variable velocities are due to the 5-

spot well pattern operation. Travel time statistics are summarized in Table 5.5. Along the direct 

path from the injection well to the pumping well, particle travel time is slightly less than 1 week. 

However, the mean particle travel time is nearly 4 weeks for the high gradient (i.e., 0.005) case 

and ~6.5 weeks for the low gradient (i.e., 0.0005) case. The longest particle capture times, ~3 

years or more, were associated with the low gradient case. While the spatial delineation of the 

capture zone pore volume in Figure 5.6 is definitive, the injection of one pore volume will 

definitely not access the entire capture zone. This is because the mass flux of injectate will be 
5-13

location-dependent and there are a few particle trajectories that will take well over two orders of 

magnitude more time to be captured than the fastest particles. This will become even more 

significant when the time scales of reaction are similar to some of the time scales of transport. 

Spatially dependent electron donor delivery will lead to preferential depletion of solid phase 

electron acceptors. 

Table 5.5 Particle travel time statistics.

Case

Particles 

Recovered 

/ Particles 

Released

Particle Travel Time in Days

Minimum Median Mean Maximum

Grid 1 Grad 0.0005 E-W 240/240 6.7 d 10.2 d 46.0 d 1250 d

Grid 1 Grad 0.005 E-W 234/240 6.6 10.0 25.5 351

Grid 1 Grad 0.0005 SW-NE 240/240 6.7 10.2 46.0 1050

Grid 1 Grad 0.005 SW-NE 234/240 6.6 10.0 27.3 556

Grid 2 Grad 0.0005 E-W 240/240 6.7 10.2 47.4 1280

Grid 2 Grad 0.005 E-W 234/240 6.6 10.0 25.3 337

Grid 2 Grad 0.0005 SW-NE 240/240 6.7 10.2 46.5 1080

Grid 2 Grad 0.005 SW-NE 234/240 6.6 10.0 27.2 548
5-14

 

Figure 5.7 Travel time histograms for particles released from the perimeter of the injection 

wells. The travel time intervals are in days: < 10, 10 to 30, 30 to 100, 100 to 300, 300 to 1000, 

and > 1000. Results are presented for the four combinations of flow(east to west, and 

southwest to northeast) and regional groundwater gradient (0.0005, 0.005).

 The second method of evaluating well field performance in the model was to consider 

transport of an inert tracer that is continuously released from each injection well. 

Concentrations at successive times up to 1000 days are contoured in Figures 5.8-5.14. A 

characteristic cloverleaf pattern forms, where concentrations are naturally greatest near the 

injection wells, and least in the four areas where flow is able to go directly from the distal aquifer 

to the extraction well. The area containing tracer grows over time, more rapidly in early time 

and more slowly later, but never stops increasing. This growth occurs even in simulations 

where diffusion and dispersion are set to zero because of numerical dispersion. Although 

numerical dispersion is caused by the discrete rather than continuous nature of the model grid 

and is a type of model error, it can be regarded as one approximation of the dispersion that 

happens in the real system. The real system will have a more complex variation in flowpaths 

and travel times that are caused by the natural variability in hydraulic conductivity and porosity. 

Figures 5.8 and 5.9 show the minor effect of grid choice on tracer concentrations, 

particularly near the well field. Grid 2 involves a larger number of smaller cells near the well 

field, and results in somewhat sharper plume edges visible in later time. However, the patterns 

are essentially the same. 
5-15

Figure 5.8 Grid 1, Gradient 5e-04 E-W, D=0, tracer concentrations over time. Green dots are 

well locations.
5-16

Figure 5.9 Grid 2, Gradient 5e-04 E-W, D=0, tracer concentrations over time. 

In addition to the cases where diffusion and dispersivity were both set to zero (D=0), cases 

were also run with diffusion coefficient = 2.5e-5 cm2

/sec (molecular diffusion in water), 

longitudinal dispersivity = 10.0 m, and transverse dispersivity = 1.0 m (D=10). This magnitude 

of dispersivity is somewhat large though not uncommon in nature, and in theory the numerical 

model could accommodate dispersivity up to 10% of the domain extent because of the choice of 
5-17

uniform material property distribution across the domain. Figure 5.10 shows the same case as 

Figure 5.9, except for having D=10. The tracer is much more spread out and, in this case, the 

lower grid resolution outside the primary zone five-spot well pattern is creating some angular 

features in the contours. 

Figure 5.10 Grid 2, Gradient 5e-04 E-W, D=10, tracer concentrations over time.

Cases with the steeper background gradient and flow from southwest to northeast are shown in 

Figures 5.11-5.13. It is evident in Figure 5.13 that the stronger regional flow field leads to 

greater skewing of the cloverleaf pattern in the downgradient direction, and some tracer reaches 
5-18

the model boundaries even with D=0. This case is stressing the ability of the grid 1 specification 

to capture the central behavior without numerical artifacts. 

Figure 5.11 Grid 1, Gradient 0.005 SW-NE, D=0, tracer concentrations over time.
5-19

For the remaining tests only the grid 2 results are presented. Figure 5.12 is the grid 2 version 

of the previous 0.005 SW-NE gradient with the D=0 specification. This clearly demonstrates 

that for the higher regional flow cases, using a larger grid can keep the tracer plume in a region 

where resolution is adequate. 

Figure 5.12 Grid 2, Gradient 0.005 SW-NE, D=0, tracer concentrations over time.
5-20

The high SW-NE gradient, high dispersion case (Figure 5.13) shows significant downgradient 

migration away from the capture zone of the central pumping well. 

Figure 5.13 Grid 2, Gradient 0.005 SW-NE, D=10, tracer concentrations over time.
5-21

The final method for evaluating the well field was to track particles initially placed throughout the 

model domain, and delineate the zone where particles were captured by the extraction well. 

Here the goal was not to evaluate the fate of injected water, but rather to see the relationship 

between the area of advection towards the extraction well, and the area of tracer presence 

caused by dispersion as well as advection. If the advection capture zone does not contain the 

tracer plume, then tracer will almost certainly be lost from the well field. Figure 5.14 shows the 

capture zone as a contour line overlaid on the 1000 day tracer concentrations. The most 

favorable cases with respect to tracer capture are those with the lower background gradient and 

D=0. However, even in those two cases, there is a portion of the tracer plume at high 

concentrations that lies outside the advective capture zone. Increasing the background gradient 

or dispersivity decreases the amount of capture, and indicates that more tracer will escape the 

extraction well. 
5-22

Figure 5.14 Tracer concentrations at 1000 days for 4 base cases (D=0) and 4 dispersion cases 

(D=10), grid 2. Black lines delineate capture zones, wherein particles within contour line are 

captured by the extraction well in advective transport.
5-23

5.3 Findings

The general findings from the field-scale simulations are the following:

• In the comparison of the steady-state and transient flow modeling approaches, it was 

found that most of the change in the flow field occurred in the first two days of the fivespot well operation. All changes in groundwater velocities were zero within 200 days. 

Although the analyses used transient velocity fields, these results justify steady-state 

flow fields for scoping and scenario comparisons.

• Flow modeling on both the small and large grids was numerically consistent with the 

linear PDEs being solved and satisfied the principle of superposition. For the flow 

phenomena of interest, the smaller grid (smaller domain, lower resolution, and fewer 

grid cells) performed very similarly to the larger more detailed grid. The larger grid was 

necessary to resolve the transport of a conservative tracer when the regional gradient 

or the dispersivity was high. 

• Operation of the 5-spot well pattern results in a nonuniform flow field even though the 

material properties are homogeneous. Travel time for particles released near the 

injection well ranged from less than a week to 3.5 years. Mean particle travel time is 

nearly 4 weeks for the high regional gradient (i.e., 0.005) case and ~6.5 weeks for the 

low regional gradient (i.e., 0.0005) case. This has implications for characterizing 

treatment in terms of injected pore volumes. While the pore volume of the capture 

zone is definitive, the time it takes to access the entire pore volume is much longer than 

the time it takes to inject one pore volume. This is because transport in the low flow 

rate regions (i.e., away from the axis between the pumping and injection wells) is 

slower and exposed to less injectate than other regions. This variability in advective 

travel time from the injection well to the pumping well results in a range of residence 

times within the capture zone that will affect the extent of kinetically controlled 

reactions. 

• While most of the material injected in the perimeter wells of the five-spot pattern will be 

captured by the central pumping well, higher regional gradients increase the 

opportunity for excursions of injectate beyond the capture zone. This will be 

exacerbated by heterogeneities in the hydraulic conductivity that alter the idealized flow 

paths examined in this study.

6-1 

6 Field-Scale Post-ISR Biorestoration Modeling 

6.1 Biorestoration Concepts

The general approach for the biorestoration of uranium ISR sites is to return the roll front to

the low pre-operational redox potential that maintained the baseline conditions. The 

fundamental assumption is that the introduction of lixiviant and duration of ISR operations has 

not irreversibly altered the mineral assemblages in a way that would prevent restoration of the 

low redox potential. In this respect, the solid phases are the largest and most important 

reservoirs for redox capacity in the subsurface. For the biorestoration technology to succeed, 

the post-ISR sediments must be returned to a redox potential that is sufficient to restore and 

sustain the baseline redox conditions. 

The objective of the field-scale simulations is to gain insight on the potential for ISR 

biorestoration and issues of performance. We use the results of the modeling of the column 

experiments in Chapter 4 to inform the field-scale modeling approach. The column experiments 

and associated modeling are not considered substitutes for well- instrumented field experiments 

that address site-specific conditions. Consequently, these hypothetical simulations of the field 

implementation of lactate biorestoration are essentially scoping analyses to identify insights and 

potential issues with the field implementation. 

 As mentioned in Chapter 4 (modeling of the column experiments), the poorly consolidated 

sand and silt dominated sediments that were recovered from the Kingsville Dome drilling were 

used directly without having to exclude larger size fractions. The column diameter was 2.22 cm 

which was appropriate for the sediments, which were < 1 mm. Ostensibly, the use of the full 

field particle size distribution eliminates or at least minimizes the need to upscale the reactive 

phases that are typically associated with smaller size fractions. For this reason, no upscaling of 

the reactivity is performed for the field scale simulations. 

In U.S. ISR operations, oxygen is commonly used in the injected lixiviant. The oxidation of 

reduced minerals in the sandstone aquifer sediment is presumed to be a key factor in the longer 

time scales associated with post-ISR restoration. The re-establishment of the low redox 

potential baseline is based on diminishing the effects of oxidants in the system using 

biostimulation of indigenous organisms to catalyze the reduction of key terminal electron 

acceptors such as U(VI). 

Sparging with oxygen gas would increase the electron donor requirements by the ratio of 

oxygen partial pressures. At ~100 m below the water table, the hydrostatic water pressure is 

nearly 10 atm. This will increase the dissolved oxygen saturation capacity by an order of 

magnitude. Using 25° C Henry’s Law coefficients for oxygen with atmospheric gas composition, 

the dissolved oxygen concentration could be 2.6 mM. Based on the stoichiometry of lactate and 

acetate oxidation coupled with oxygen reduction, about 2 moles of oxygen can be reduced for 

each mole of lactate reacted. Thus, a little more than 1 mM lactate will completely consume 

the accessible oxygen assuming that dissolved oxygen is at saturation at end of the ISR 

process. The 10 mM lactate injectate should stimulate sufficient bioreduction to deplete the 

oxygen. Thus, the expectation is that the groundwater brought to the surface by the production 
6-2 

well will be anaerobic after bioremediation is implemented. In fact, oxygen in groundwater 

sampled about two years after ISR processes stopped was < 0.015 mM, Table 3.1, suggesting 

that this evaluation is very conservative. 

In the handling of the produced groundwater at the surface before re-injection, a potentially 

important consideration is the maintenance of the anaerobic conditions. The issue is whether

greater than suboxic levels of oxygen adversely affect the microorganisms that accomplish the 

intended bioreduction of terminal electron acceptors. It also may be important to keep the 

oxygen levels as low as possible to drive the diffusion of oxygen out of intra- and inter-particle 

zones that have limited transport access to the bulk groundwater flow. Anaerobic groundwater 

will also deplete oxygen from entrapped gas bubbles through partitioning with the aqueous 

phase. If oxygen sparging was used during the ISR process, exsolution of oxygen as the 

pumped groundwater reaches the ground surface is possible but only in the initial pore 

volume(s). The electron donor should stimulate microorganisms that will use dissolved oxygen 

as a terminal electron acceptor, resulting in the conversion of oxygen to water. 

6.2 Modeling Approach

The specification of the field-scale simulation is based on the reaction network, initial and 

boundary condition geochemistry developed in the Chapter 4 modeling of the column 

experiments and the Chapter 5 flow and transport studies. The biorestoration simulation 

assumes the 0.0005 west to east regional gradient case where the flow field is aligned with the 

grid (i.e., not the southwest to northeast flow field that was diagonal to the grid). The initial and 

upgradient influent boundary condition chemistry for the regional system is the boundary 

condition chemistry in Table 4.4 without lactate. The solid phases are based on Table 4.2. 

The assumption here is that the production well in the center of the five-spot pattern will be 

initially operated alone as a groundwater sweep. This can be run until the initial transient of 

elevated metals concentrations decreases and begins to approach a steady state. Typically, 

this can require 1-2 pore volumes. After the groundwater sweep, the biostimulation will be 

initiated by introducing groundwater amended with 10 mM lactate in each of the 4 perimeter 

wells of the 5-spot pattern at the 10 gpm groundwater injection rate. The assumption is that the 

injected groundwater is from a background well with no elevated metals concentrations. In this 

case, we assume a baseline uranium concentration of 0.075 mg/L (0.315 μM). All other 

background groundwater components are as listed in Table 4.4. The pumping rate for the 

central production well in the 5-spot pattern can be operated at the summation of the injection 

rates at the 4 wells. Unlike the ISR operation where excursions of lixiviant and metals cannot 

be tolerated, minor injectate excursions of lactate are not considered a threat to the water 

quality. Consequently, the produced water can all be recycled, provided the water quality is 

satisfactory. If treatment is necessary, standard ion exchange and reverse osmosis techniques 

described in Chapter 2 can be employed. 

Two injectate formulations are modeled. The first case, presented in Section 6.3, uses 

lactate amendment in the injectate under the assumption that recycling of bioreduced water 

from the production well effectively removes the U(VI) and sulfate terminal electron acceptors. 
6-3 

The second case, presented in Section 6.4, uses injectate based on lactate amendment of 

groundwater from a background well with the terminal electron acceptors present.

6.3 Field-Scale Biorestoration Modeling Results

As in the modeling of the column experiment, the injected lactate is consumed in two 

microbially-mediated TEAP reactions, phyllosilicate Fe(III) and sulfate, whereas the acetate 

produced by these reactions is consumed in TEAP reactions involving crystalline (i.e., goethite) 

and phyllosilicate forms of Fe(III) as well as aqueous U(VI). The solid phase phyllosilicate Fe(III) 

is immobile with 98.5% of the biogenic iron remaining in the clay structure. Note that aqueous 

phase components will be physically displaced and mixed by the injected water. Relative to the 

column experiment, the magnitudes and time scales of the TEAP reaction products are 

profoundly affected by the pumping rates and well configuration. For example, the rate of 

electron donor delivery to the sediments near the well bore is three orders of magnitude more 

than the column experiment. Consequently, a parcel of injectate will react with considerably 

more than 10 cm of sediment and sediments near the well bore will be exposed to 3 orders of 

magnitude more electron donor. 

After 2 days of biostimulation, lactate is largely unreacted and within a 5 m radius of the 

injection wells (Figure 6.1). From Table 5.5, the median residence time is on the order of 10 

days. Thus, the bulk of the injectate is still in the vicinity of the injection well. At this point, the 

pumping well has only a minor impact on the geometry of the lactate distribution. The 

outermost fringe of the lactate distribution, which contains the lactate that has been in the 

system the longest is approximately half of the initial lactate concentration. While lactate is 

being consumed in TEAP reactions, dispersion also contributes to the lower concentrations on 

the outer fringe of the injectate. 
6-4 

Figure 6.1 Lactate concentration (M) distribution after 2 days of injection from perimeter wells. 

After 2 days, the acetate concentrations near the injection well are essentially zero 

corresponding to where the lactate reaction kinetics are slow relative to the high transport rates 

near the injection well (Figure 6.2). The highest acetate concentrations, ~3 mM, are in a central 

band of the injectate fringe. Lower concentrations are found inside and outside of this band. 

This represents the radius where the transport time scales and TEAP kinetics are of the same 

order. In this case, the lower acetate concentrations just inside the ring of peak concentrations 

are due to less lactate being oxidized. Conversely, outside this ring, the lower concentrations 

are due to consumption of acetate by Fe(III) TEAP reactions and the dispersive transition 

across the acetate front. Slightly more skewing in the geometry of the acetate distributions is 

due to the pumping well drawing injectate towards the middle of the 5-spot well pattern.
6-5 

Figure 6.2 Acetate concentration (M) distribution after 2 days of injection from perimeter wells. 

Aqueous Fe(II) concentrations are very similar in character to the acetate concentrations 

with a maximum ~1 mM concentration (Figure 6.3). This implies that the acetate-driven Fe(III) 

bioreduction is controlling the dissolved Fe(II) behavior. In the model, lactate participates in the 

phyllosilicate Fe(III) TEAP reaction but not the crystalline Fe(III) TEAP reaction. This is an 

important distinction since most of the biogenic Fe(II) from the reduction of phyllosilicate Fe(III) 

is reduced in place (i.e., not liberated into solution); whereas the bioreductive dissolution of 

crystalline Fe(III) (i.e., FeOOH), which is only associated with acetate electron donor, liberates 

all bioreduced iron into solution. 
6-6 

Figure 6.3 Aqueous Fe(II) concentration (M) distribution after 2 days of injection from perimeter 

wells. 

Figure 6.4 illustrates the situation if we assume sulfate is progressively removed from 

solution by reinjecting and refortifying the injectate with lactate. In this case, the low sulfate 

injectate is displacing the background sulfate resulting in very low concentrations near the well. 

On the fringes of the impacted zone, sulfate concentrations between the ambient and injectate 

concentrations are the result of dispersion and sulfate bioreduction.
6-7 

Figure 6.4 Sulfate concentration (M) distribution after 2 days of injection from perimeter wells.

The aqueous U(VI) concentration distribution after 2 days of lactate biostimulation displays 

complex behavior (Figure 6.5). While aqueous U(VI) is being displaced by the injectate, the low 

bicarbonate injectate is also favoring less aqueous uranium complexation and more uranium 

surface complexation. This results in more adsorbed U(VI) in the presence of the injectate as it 

moves outward. Within the outer fringe of the injectate migration, we see a ring of elevated 

aqueous U(VI) concentrations, ~5 μM, which is still lower than the ambient 20 μM 

concentrations. On either side of this ring are lower U(VI) concentrations. The ring represents 

the point where the displaced aqueous U(VI) has been pushed out. The lower concentrations 

on the outside of the ring are the result of U(VI) TEAP bioreduction reactions with acetate from

the oxidation of lactate. In this case, the kinetics of lactate consumption are sufficiently slow to 

limit the highest acetate to injectate that has been in the aquifer the longest, i.e., the fringe 

extent of the injectate migration. 
6-8 

Figure 6.5 Aqueous U(VI) concentration (M) distribution after 2 days of injection from perimeter 

wells.

6.3.1 10 Days

The modeling results after 10 days of lactate biostimulation from the four corner injection 

wells show that injectate has just arrived at the central production well. This corresponds with 

the mean advection travel time from the injection wells to the pumping well. Aqueous U(VI) 

initially in each injectate lobe is being displaced by the injectate. The bulk of the U(VI) is initially

adsorbed to the sediments. In Figure 6.6, adsorbed U(VI) concentrations are decreasing due to 

desorption in response to the displacement of the aqueous U(VI). 
6-9 

Figure 6.6 Adsorbed U(VI) concentration (M) distribution after 10 days of biostimulation from 

injection wells.

The bioreduction of aqueous U(VI) in the acetate-based TEAP reaction is also taking place. 

Thus, desorbed U(VI) is decreased subject to the availability of acetate and the U(VI) TEAP 

reaction kinetics. Figure 6.7 shows that the areas of highest desorption are correlated with the 

areas of highest precipitation of U(IV) mineral. Once again, the lactate oxidation kinetics 

controlling acetate production are sufficiently slow relative to the near-injection well transport 

rates that acetate-based TEAP reactions (e.g., FeOOH and U(VI)) are not significant in the 

vicinity of the injection well. The highest U(IV) concentrations occur intermediate between the 

injection well and the lobe edges. This is where the displaced U(VI) is most available. U(IV) 

concentrations drop off near the lobe edges with the depletion of aqueous U(VI) and acetate as 

well as dispersive mixing across the injectate front. The implication here is that the interplay of 

transport, desorption, and U(VI) bioreduction will dictate how much aqueous U(VI) reaches the 

pumping well. 
6-10

Figure 6.7 U(IV) concentration (M) distribution after 10 days of biostimulation from injection 

wells.

The bulk of the Fe(II) in solution after 10 days originated in the FeOOH (goethite) crystalline 

Fe(III) pool. With the exception of the near-injection well zone where little acetate was 

produced, FeOOH is depleted throughout the injectate impacted zone (Figure 6.8). 
6-11

Figure 6.8 FeOOH (goethite) concentration (M) distribution after 10 days of biostimulation from 

injection wells.

The production of aqueous Fe(II) from the bioreductive dissolution of FeOOH is not 

apparent from the distribution of aqueous Fe(II) concentrations. Most illuminating is the 

distribution of FeS, which is similar in character to the FeOOH (Figure 6.9). This implies that 

sufficient sulfide is available to react with the aqueous Fe(II) in the injectate lobe(s) to form FeS.
6-12

Figure 6.9 FeS concentration (M) distribution after 10 days of biostimulation from injection wells.

Another potential solid phase sink for Fe(II) is through adsorption. In this case, relatively 

small amounts of Fe(II) are sorbed along a narrow fringe outside of the areas of primary FeOOH 

dissolution and FeS precipitation (Figure 6.10).
6-13

Figure 6.10 Adsorbed Fe(II) concentration (M) distribution after 10 days of biostimulation from 

injection wells.

The lactate-based phyllosilicate Fe(III) consumption is approximately 3e-4 M over the 

injectate lobe(s) (Figure 6.11). This relatively small change implies that the bulk of lactate 

oxidation is due to the sulfate TEAP. In part, this is due to the reaction stoichiometry where 

almost 3 moles of lactate are oxidized to reduce 1 mole of sulfate. Thus, the approximately 

1.5e-3 M increase in FeS concentration across the injectate lobes results from the reduction of 

1.5e-3 M sulfate in which 4.5e-3 M lactate was consumed. 
6-14

Figure 6.11 Lactate-driven phyllosilicate Fe(III) concentration (M) distribution after 10 days of 

biostimulation from injection wells.

The reduction of phyllosilicate Fe(III) in the injectate lobe via acetate-based TEAP reaction 

is approximately 1.5e-2 M after 10 days of lactate biostimulation (Figure 6.12). This compares 

to 5e-3 M of FeOOH. Thus, the bulk of acetate is consumed in the phyllosilicate Fe(III) TEAP 

reaction. This is consistent with the reaction rate, which is faster than for FeOOH.
6-15

Figure 6.12 Acetate-driven phyllosilicate Fe(III) concentration (M) distribution after 10 days of 

biostimulation from injection wells.

 

After 10 days of lactate biostimulation, aqueous sulfide is found primarily in a narrow zone 

inside the outer edge of the injectate lobe (Figure 6.13). The relative absence of aqueous 

sulfide in the lobe interior is due, in part, to removal from solution via FeS mineral precipitation 

(Figure 6.9). As indicated by the sorbed Fe(II) distribution (Figure 6.10), there is a fringe of 

Fe(II) outside of the aqueous sulfide edge. At this location, lactate has been consumed, which 

precludes sulfide production. Acetate and Fe(III), however, are present resulting in the 

production of Fe(II) from the bioreduction of Fe(III) minerals. 
6-16

Figure 6.13 Aqueous sulfide concentration (M) distribution after 10 days of biostimulation from 

injection wells.

6.3.2 30 Days

As expected from the tracer studies, lactate from the injection wells reaches the production 

well in ~10 days. The lactate concentration contours in Figure 6.14 at 30 days of biostimulation 

show that lactate arriving at the well has generally undergone minimal attenuation. Only the 

lactate transported along the flow paths with the longest travel times (i.e., outer edges of the 

four injectate plumes) exhibit attenuation. This is due to a combination of lactate oxidation and 

dispersion across the edges of the plume.
6-17

Figure 6.14 Lactate concentration (M) distribution after 30 days of injection from perimeter wells.

Figure 6.15 shows that after 30 days acetate is found primarily along the perimeter of each 

injectate lobe. While acetate is being generated wherever lactate and electron acceptors are 

available, the kinetics of the lactate oxidation are sufficiently slow relative to the transport rates 

near the injection well that very little acetate is generated within a radius of ~10 m from the 

injection well. Outside of that radius, lactate oxidation produces sufficient acetate to drive Fe(III) 

TEAP reactions that begin to deplete crystalline (i.e., FeOOH) and phyllosilicate Fe(III) inside 

the injectate lobes. It is that process that consumes acetate in the lobes whereas along the 

perimeter of the lobes there are slower transport pathways, allowing more time for lactate 

oxidation which produces more acetate than the consumption in the Fe(III) TEAP. As in the 

earlier result at 2 days, the highest acetate concentrations occur in a central band through the 

outer fringe of the injectate lobe. This is where lactate has had sufficient time to react to 

produce acetate but insufficient time for acetate to significantly react with Fe(III) to deplete the 

acetate concentration. Basically, this location is a balance between transport and reaction 
6-18

rates. On the outside of this acetate fringe, the lower acetate concentrations reflect the longer 

residence time, which allows more acetate to react. On the inside of this acetate fringe, the 

lower acetate concentration is the result of shorter residence time for lactate conversion to 

acetate and the consumption of acetate by the TEAP reactions.

Figure 6.15 Acetate concentration (M) distribution after 30 days of injection from perimeter 

wells.

In Figure 6.16, the elevated Fe(II) concentrations are found on the perimeter of the four 

injectate lobes, distributed similarly to the acetate concentrations. In this case, however, the 

highest concentrations are in the central band furthest from the production well. This 

corresponds to the location where lactate has been in the system for the longest duration, which 

allows more time for acetate to be produced, and more time for that acetate to react with Fe(III) 

minerals. It should be noted that this Fe(II) distribution does not necessarily reflect where the 

Fe(II) originated from. The spatial distribution of goethite depletion, which occurs over much of 

the injectate lobe (described below), indicates where Fe(II) was initially produced. The 
6-19

distribution of aqueous Fe(II) on the perimeter of the injectate lobes is the net result of the 

nonuniform flow field, lactate injection, and rates of Fe(III) and sulfate bioreduction and FeS 

mineral precipitation.

Figure 6.16 Aqueous Fe(II) concentration (M) distribution after 30 days of injection from 

perimeter wells.

Nearly all of the Fe(II) in solution is from the acetate-driven bioreduction of crystalline Fe(III) 

represented by goethite. In Figure 6.17, the zones of goethite depletion are clearly transport 

controlled. Near the well, very little acetate has been produced, limiting the Fe(II) that can be 

produced. Conversely, goethite is depleted within the injectate lobes away from the fastest flow 

paths to the production well. 
6-20

Figure 6.17 FeOOH (goethite) concentration (M) distribution after 30 days of injection from 

perimeter wells.

After the initial displacement of aqueous sulfate, there is essentially no sulfate in the 

injectate lobes (Figure 6.18). 
6-21

Figure 6.18 Sulfate concentration (M) distribution after 30 days of injection from perimeter wells.

The sulfide precipitated as FeS inside the injectate lobes after 30 days (Figure 6.19) is 

similar to the result after 10 days with a slightly larger spatial extent and larger zones of high 

concentrations. While sulfate from outside the injectate front can be entrained into the outer 

edges of the lobe by dispersion, the effect has minimal impact because of the small amount of 

sulfate that can contact and react with lactate to form sulfide. 
6-22

Figure 6.19 FeS concentration (M) distribution after 30 days of lactate biostimulation.

The sulfide concentration distribution (Figure 6.20) is largely unchanged from the 10 day 

result implying a pseudo-steady state of low-level sulfide production has set up.
6-23

Figure 6.20 Aqueous sulfide concentration (M) distribution after 30 days of lactate 

biostimulation.

Figure 6.21 shows the depletion of aqueous U(VI) throughout the injectate lobes after 30 

days. The continuous cycle of lowering aqueous U(VI) concentrations via bioreduction to U(IV) 

and driving desorption of U(VI) from the sediments has led to a three orders of magnitude 

decrease in aqueous U(VI). The negligibly small aqueous U(VI) concentrations (~1.0E-8 M) 

that are still present are due to the desorption of the adsorbed U(VI). 
6-24

Figure 6.21 Aqueous U(VI) concentration (M) distribution after 30 days of lactate biostimulation.

Figure 6.22 shows that the adsorbed U(VI) is largely depleted with the exception of the near 

well zone. As the adsorbed U(VI) controls the U(VI) in solution that is available for bioreduction, 

the U(VI) in the system after 30 days is negligible. 
6-25

Figure 6.22 Adsorbed U(VI) concentration (M) distribution after 30 days of lactate biostimulation.

With much less U(VI) available in the system, there is negligible change in the amount of 

U(IV) produced in the system from acetate-stimulated bioreduction from the 10-day result

(Figure 6.23). 
6-26

Figure 6.23 U(IV) concentration (M) distribution after 30 days of lactate biostimulation.

6.3.3 Modeling Results: Injectate without Terminal Electron Acceptors 

The presented modeling results were based on an injectate without terminal electron 

acceptors (i.e., sulfate, U(VI)). The assumption was that the injectate would be recycled and the 

terminal electron acceptors would be removed in the bioreduction process. Under these

simulated conditions, the general findings are

• A large fraction of the injected 10 mM lactate was not consumed by the time it reached 

the production well. This was due, in part, to the absence of sulfate in the injectate as 

well as the lack of a solid phase sulfate source. This limited the growth of sulfate 

reducing bacteria biomass, which is a direct control on the lactate reaction rate. 
6-27

• Despite the relatively small fraction of lactate oxidized, the acetate produced from that 

oxidation was still sufficient to drive the reduction of the available U(VI). This implies 

that smaller lactate concentrations are likely to be equally effective at U(VI) bioreduction. 

• Aqueous U(VI) will be displaced by the injectate and where bicarbonate from the 

microbially-mediated oxidation of lactate is sufficiently high, desorption of adsorbed 

U(VI) results. This process rapidly depletes the reservoir of labile U(VI) residing on the 

sediments. 

• There is very little change in the mass of U(IV) precipitated after 30 days despite 

continuing, albeit rapidly declining, desorption from the sediments.

6.4 Background Injectate 

In the previous section, the modeling assumed that the biologically-mediated TEAP 

reactions induced by the injection of 10 mM lactate would remove sulfate and U(VI) from the 

recycled groundwater used for the injectate. In this section, the modeling examines the other 

end member where the introduced injectate continuously retains the background water quality, 

including 10.6 mM sulfate and 0.3 uM U(VI), (assumed baseline concentration) augmented by 3 

mM lactate. The lowering of the injected lactate concentration from the 10 mM used previously 

to 3 mM reflects the assertion that smaller concentrations would still be effective in this modeled 

field scenario.

6.4.1 Lactate

In Figure 6.24, the continuous inclusion of the background terminal electron acceptors and 

lower 3 mM lactate concentration in the injectate leads to a rapid quasi-steady state of lactate 

concentrations limited to a small zone near the well. The distributions at days 2, 10, 20, and 30 

are essentially identical. In this case, the addition of background levels of sulfate has resulted in 

the growth of SRB biomass which the lactate reaction rate is dependent upon. The quasisteady state balances the lactate supply with the consumption through the lactate-based TEAP 

reactions. 
6-28

Figure 6.24 Lactate concentration distribution using continuous background concentrations of 

electron acceptors at 2, 10, 20, and 30 days.

6.4.2 Acetate

While the lactate concentration distribution is largely unchanged, the continuous production 

of acetate from the microbially-mediated lactate oxidation reactions results in a steadily 

increasing zone of elevated acetate concentrations (Figure 6.25). The growth of the elevated 

acetate zone is asymptotic over time as the highest acetate concentrations never reach the 

production well.

Figure 6.25 Acetate concentration distribution using continuous background concentrations of 

electron acceptors at 2, 10, 20, and 30 days.

6.4.3 U(VI)

While there is a low U(VI) background concentration in the injectate, 0.3 uM, the simulated 

aqueous U(VI) concentration distribution (Figure 6.26) is very similar to the previous injectate 

simulations. Aqueous U(VI) is largely displaced and is minimally supplied with U(VI) desorbing 

from the sediments.

2 Days 10 Days 20 Days 30 Days

2 Days 10 Days 20 Days 30 Days
6-29

Figure 6.26 U(VI) concentration (M) distribution at 2, 10, 20, and 30 days using injectate with 

continuous background concentrations of electron acceptors.

6.4.4 Adsorbed U(VI)

Similar to the previous injectate case, the adsorbed U(VI) is being largely depleted over the 

first 30 days (Figure 6.27). As the adsorbed U(VI) is the principal pool of labile U(VI) available 

for bioreduction, its depletion limits the amount of U(IV) produced. 

Figure 6.27 Adsorbed U(VI) concentration distribution at 2, 10, 20, and 30 days using injectate 

with continuous background concentrations of electron acceptors.

2 Days 10 Days 20 Days 30 Days

2 Days 10 Days 20 Days 30 Days
6-30

6.4.5 U(IV)

The production of U(IV) is thus asymptotic and the distribution of U(IV) concentrations is 

generally unchanged from 10 to 30 days (Figure 6.28). 

Figure 6.28 U(IV) mineral concentration (M) distribution at 2, 10, 20, and 30 days using injectate 

with continuous background concentrations of electron acceptors.

6.4.6 Fe(II)

Fe(III) is the principal terminal electron acceptor based on the amount consumed in 

microbially mediated TEAP reactions. Figure 6.29 shows the asymptotic growth of Fe(II) in 

solution over the first 30 days. Highest concentrations are generally along the edges of the 

injectate lobes. The low aqueous Fe(II) concentrations in the vicinity of the well is the result of a 

combination of processes. The most reacted pool of Fe(III), FeOOH, is acetate-based. This 

means the availability of the acetate that is fueling the crystalline Fe(III) bioreduction is subject 

to lactate oxidation kinetics. Furthermore, the production of sulfide via lactate-based 

bioreduction of sulfate provides sulfide to react with the Fe(II) produced near the well. 

 

Figure 6.29 Aqueous Fe(II) concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors.

2 Days 10 Days 20 Days 30 Days

2 Days 10 Days 20 Days 30 Days
6-31

6.4.7 Phyllosilicate Fe(III)

Figures 6.30 and 6.31 show the minimal amount of phyllosilicate Fe(III) that is converted by 

either lactate- or acetate-based bioreduction. 

Figure 6.30 Lactate-based phyllosilicate Fe(III) concentration distribution at 2, 10, 20, and 30 

days using injectate with continuous background concentrations of electron acceptors.

Figure 6.31 Acetate-based phyllosilicate Fe(III) concentration distribution at 2, 10, 20, and 30

days using injectate with continuous background concentrations of electron acceptors.

6.4.8 Goethite

Figure 6.32 shows the depletion of FeOOH mineral over the first 30 days of simulation. 

Only the small zone nearest the injection well is largely unreacted. The reductive FeOOH 

dissolution is controlled by acetate production which increases with the increasing residence 

time afforded by the slowing of pore velocity away from the wells. 

2 Days 10 Days 20 Days 30 Days

2 Days 10 Days 20 Days 30 Days

2 Days 10 Days 20 Days 30 Days
6-32

Figure 6.32 FeOOH (goethite) concentration distribution at 2, 10, 20, and 30 days using 

injectate with continuous background concentrations of electron acceptors.

6.4.9 FeS

The production of Fe(II) in the center of the lobe due to FeOOH dissolution is not reflected in 

the aqueous Fe(II) concentration distribution. This is because the formation of FeS (Figure 

6.33) is consuming most of the aqueous Fe(II) in this region. In fact, the highest concentrations 

of FeS are in a zone centered around the injection well. 

Figure 6.33 FeS mineral concentration distribution at 2, 10, 20, and 30 days using injectate with 

continuous background concentrations of electron acceptors.

6.4.10 Sulfate

While background sulfate is continuously being introduced through the injection well, sulfate 

concentrations are fairly uniform across the injectate lobe with the exception of a small zone 

around the injection well (Figure 6.34). This indicates that most of the 1 to 2 mM sulfate 

attenuation has occurred near the injection well and that these lower sulfate concentrations are 

then transported with negligible reaction after that. This is consistent with the rapid lactate 

consumption near the injection well which is primarily driven by the sulfate TEAP reaction, as 

evidenced by the negligible amount of lactate-based phyllosilicate Fe(III) reaction. 

2 Days 10 Days 20 Days 30 Days
6-33

Figure 6.34 Aqueous sulfate concentration distribution at 2, 10, 20, and 30 days using injectate 

with continuous background concentrations of electron acceptors.

6.4.11 Sulfide

Unlike the residual sulfate that is transported with negligible reaction, the sulfide produced 

near the well reacts with the available Fe(II) to form FeS mineral. This occurs until the sulfide is 

largely depleted (Figure 6.35). This results in aqueous Fe(II) being found only in the parts of 

the injectate flow field that have been in the system the longest, the perimeter of the injectate 

lobes. 

Figure 6.35 Aqueous sulfide concentration distribution at 2, 10, 20, and 30 days using injectate 

with continuous background concentrations of electron acceptors.

 

2 Days 10 Days 20 Days 30 Days

2 Days 10 Days 20 Days 30 Days
6-34

6.5 Summary of Field-Scale Model Findings

The findings that follow are based on a hypothetical scoping simulation of field-scale 

behaviors. The conceptualization of modeled processes, properties, and conditions is based on 

laboratory studies and general attributes of uranium ISR sites. 

Two end member injectate water chemistry scenarios were modeled. In Section 6.3, the 

simulated injectate assumed that the recycling of biorestored groundwater with lactate 

amendment would effectively remove sulfate and U(VI) from solution, so these electron 

acceptors were not included in the injectate. In Section 6.4 the simulated injectate included 

background sulfate and U(VI) concentrations. Under both injectate scenarios, the aqueous 

uranium concentrations rapidly decreased to well below pre-baseline concentrations. This 

occurred after 30 days of lactate biostimulation using the reactions and rates determined from 

the laboratory studies. In the injectate-affected zones (“lobes”), aqueous and adsorbed U(VI) 

was generally depleted and U(IV) precipitated. The simulations were terminated after 30 

simulated days because of the depletion of the solid phase U(VI) source, which led to 

essentially unchanging uranium concentrations in aqueous, sorbed, and mineral forms. 

The behavior of Fe and S, however, was much more sensitive to injectate composition. In 

the absence of the electron acceptors, FeS (Figure 6.19) and aqueous sulfide (Figure 6.20) are 

distributed primarily away from the injection wells. This is consistent with displacement of 

aqueous sulfate by the injectate coupled with rate-limited sulfate bioreduction at the 

displacement front where lactate and sulfate coexist. Under the scenario with terminal electron 

acceptors in the injectate, the injection of sulfate and lactate allows continuous sulfate 

bioreduction in the vicinity of the well (Figure 6.35). Aqueous Fe(II), also being continuously 

produced principally by goethite bioreduction, reacts with the sulfide to form the secondary 

mineral FeS (Figure 6.33). This local accumulation of FeS near each injection well has about 

six times higher concentrations than the scenario without sulfate in the injectate. This could 

lead to potentially adverse effects on the hydraulics of the wellbore environment and thus, 

should be considered in the design controls for lactate, sulfate, and iron in the injectate.

 While the time scale of uranium bioremediation in these simulations is similar to behaviors 

observed in bioremediation studies in less contaminated shallow aquifer systems, the 

idealizations in the modeling specification are thought to play a significant role in these results. 

In particular, the assumption of uniform material properties and the general accessibility of the 

solid-associated uranium for reaction will lead to the most rapid and complete uranium 

bioreduction. Higher initial concentrations of solid-associated U(VI), lower permeability zones 

that result in a slower release of U(VI), and more kinetic limitation of reaction rates would 

require longer periods of bioremediation to achieve similar results. While the modeling can 

accommodate more realistic attributes, it would still not be a substitute for a well-conceived 

field-driven experimental study. In part, this is due to the sensitivity of the engineered behaviors 

to site-specific geology, hydrology, biology, and geochemistry. In the absence of actual site 

geometries, conditions, and properties, the simulations are still capable of identifying issues for 

the biorestoration of uranium ISR sites as well as the design of a well-instrumented field study. 

• The bioreduction of aqueous U(VI) to insoluble U(IV) removes U(VI) from the water 

column driving U(VI) desorption from the sediments. This cycle of liberating U(VI) into 
6-35

solution followed by bioreduction progressively depletes the U(VI) from the treatment 

zone and establishes a lower redox potential for the system. Under the idealized 

modeling assumptions, the time scale of system stabilization is on the order of months. 

• Even in the idealized hydrology and homogeneous property distribution, there are 

spatially variable distributions of uranium and other components of interest. This is 

because of the interplay between the spatially variable flow field (from the 5-spot well 

pattern) and the reaction kinetics which controls the availability of reactants for primary 

and subsidiary reactions. This results in behaviors that are different from the laboratory 

experiments even though the same reaction network and similar concentrations are 

employed: 1) rapid and complete consumption of lactate via the sulfate TEAP reaction, 

and 2) negligible impact of phyllosilicate Fe(III) TEAP reactions (lactate- and acetatebased). 

• Injectate water quality can significantly affect the reaction rates of electron donors. This 

has implications for the recycling of groundwater from the production well as injectate. 

Assuming the maintenance of anaerobic conditions of the produced groundwater, 

terminal electron acceptors such as sulfate and U(VI) will be progressively removed. In 

particular, the progressive removal of sulfate as produced groundwater is recycled, 

amended with lactate, and re-injected, should reduce the amount of lactate needed. 

Thus, it’s possible that the lactate supply can be reduced based on the monitoring of the 

produced groundwater. While this might reduce the potential for biofouling, it will also 

decrease the amount of FeS produced, which may slow the return to and/or weaken the 

maintenance of a stabilized low redox potential.

• In general, lactate concentrations considerably less than the 10 mM used in the 

laboratory column experiments are shown to be equally effective for U(VI) bioreduction. 

In an actual field implementation, this will be heavily dependent on the in situ rates.

7-1 

7 Conclusions

A common restoration experience at uranium ISR sites is the persistence of elevated

dissolved U(VI) (and other redox-sensitive metals) in spite of groundwater sweep, and above

ground treatment (e.g., ion exchange, reverse osmosis). Restoring groundwater after uranium 

ISR extraction to the pre-operational baseline condition is not readily achievable. It typically 

takes many years to restore uranium ISR sites to regulatory standards. One motivation for this

study was recent research, primarily in shallow aquifers, demonstrating that indigenous 

microorganisms could be stimulated to catalyze the conversion of U(VI) in contaminated 

groundwater to immobile U(IV), even when abiotic reductants were ineffective. In some of these 

uranium bioremediation field studies, vanadium and selenium present in the groundwater were 

also opportunistically reduced and removed from solution (Williams et al., 2012; Yelton et al., 

2012). 

The principal conclusion from this study is that engineered biorestoration of post-uranium 

ISR sites is a potentially viable alternative to standard restoration approaches. The 

experimental and modeling studies presented here suggest that biostimulation of the indigenous 

bacterial populations can be effective in lowering aqueous concentrations of uranium to 

acceptable levels. This takes place despite a very depauperate population of bacteria that may 

be the result of ISR operations or may be a natural characteristic of the site. 

The naturally reducing and oxygen-depleted conditions that exist in uranium roll front aquifer 

systems are attributed, in part, to anaerobic bacteria in the presence of bioavailable organic 

carbon. These conditions are associated with very low concentrations of aqueous uranium, as 

well as redox sensitive metals (e.g., vanadium, selenium). Thus, the biorestoration principle 

proposed here is to return uranium ISR sites to their native hydrologic and geochemical 

conditions. The approach intends to (re-)establish an anaerobic microbial community that is 

operationally similar to the one that mediated the original formation of the uranium roll front ore 

zone. Working within the processes, properties, and conditions that a given aquifer will naturally 

support is thought to be more likely to succeed than engineering a new regime that must be 

artificially maintained. Conversely, the farther the targeted conditions are manipulated away 

from the natural unmitigated state of the subsurface system, the more difficult it will be to 

achieve long-term success. 

The potential for engineered biorestoration was supported by batch, column, and solid 

phase studies presented in Section 3.0 using microbiologically preserved sediments from a 

post-ISR site at the Kingsville Dome operation of Uranium Resources, Inc. One advantage of 

using the Kingsville Dome sediments was that the maximum particle size of the field sample

was less than 1 mm, which allowed the sediments to be used directly in the column experiments 

without the removal of the larger size fractions. This minimized one scale-up issue regarding the 

appropriateness of using the set of processes, reactions, and rates characterized from the 

column experiments to model the hypothetical field situation. The continuous introduction of 10 

mM lactate into sediment columns resulted in the stimulation of indigenous microorganisms that 

continuously and effectively removed 20 μM (5 mg/L) U(VI) in the influent over the ~1 day

column residence time, with resulting effluent uranium concentrations of less than 0.05 μM. 
7-2 

Analyses of sediments recovered from columns at the end of the biostimulation stage showed 

that the sequestered uranium was primarily U(IV). 

The conceptualization of the processes simulated during biostimulation begins with the 

bioreduction of aqueous U(VI) to sparingly soluble U(IV). This leads to the removal of U(VI) 

from solution and the elevation of bicarbonate concentrations, conditions that 

thermodynamically favor the desorption of U(VI) from the sediments. This cycle of U(VI) 

bioreduction and desorption progressively depletes aqueous U(VI) from the treatment zone and 

establishes a lower redox potential for the system. Under the idealized modeling assumptions, 

the time scale of system stabilization is on the order of months.

While the efficiency of uranium removal had some similarity to previous experimental 

investigations of biostimulation in shallow aquifer sediments, an important finding was that the 

biogeochemical processes also had some distinct differences: 1) lactate appeared to inhibit the 

bioreduction of U(VI) meaning that acetate, the product of lactate oxidation, was the primary 

electron donor for the U(VI) TEAP, 2) sulfate bioreduction was relatively small and independent 

of excess acetate, and 3) lactate- and acetate-specific Fe(III) pools were the most 

straightforward modeling assumptions to account for the subsequent rise in acetate 

concentrations to the near-maximum stoichiometric conversion from lactate. The interpretation 

of biogeochemical processes that are different than those observed in shallow sites, is 

strengthened by electron balances from the column studies that suggest, given the available Fe 

pools in the sediment, other processes are needed to account for the observed lactate/acetate 

consumption. 

An important observation is that despite the very low biomass in the post-ISR sediment, 

biostimulation of metal-reducing bacteria led to essentially complete conversion of U(VI) to 

immobile U(IV), as well as a lowered redox potential that maintains the stability of the reduced 

uranium. The sediment sample from the field contained sulfate reducing bacteria, but no

detectable Geobacter, which is commonly identified in shallow aquifer field biostimulation 

experiments. Lactate biostimulation of post-ISR sediments resulted in a different consortium of 

dominant microorganisms than observed in shallow aquifer sediments. This was characterized, 

in part, by the dominance of Rhizobium species that have been found in uranium-contaminated 

groundwater and sulfate reducing bacteria observed on filters collected during the late 

bioreduction phase of the column experiments. Five species of Geobacter were identified but at 

very low abundances. An open question is: “To what extent did exposure to the oxygenated 

lixiviant disrupt the pre-ISR microbial community?” 

The effectiveness of biorestoration as a viable long-term remediation strategy requires 

conditions that maintain the stability of bioreduced uranium and other solid-associated redoxsensitive contaminants. As discussed in Chapter 3, introducing suboxic levels of dissolved 

oxygen (6 μM) into the column of bioreduced sediments resulted in the oxidative remobilization 

and elution of uranium. Clearly, dissolved oxygen can disrupt biorestoration and is an important 

parameter to monitor. A more thorough understanding of mechanisms resulting in oxidation and 

remobilization of bioreduced uranium is needed.

In the field, dissolved oxygen can enter the bioreduction zone through a variety of scenarios. 

It is conceivable that the pre-operational hydrology and oxidant load in the ore zone, (i.e., the 
7-3 

natural conditions that will be relied upon to maintain the restored ISR site) may limit the 

biorestoration efficiency or effectiveness. Over time, uranium ISR operations will load oxygen 

into less mobile regions of the porous media. After uranium ISR operations cease, the residual 

oxygen in these less mobile regions may become long-term diffusion-limited sources. Even 

during restoration, handling at the ground surface of the recycled and reinjected groundwater 

may inadvertently allow oxygen to be continuously introduced into the aquifer. Another scenario 

is for oxygenated water upgradient of the uranium roll front to be drawn through the ore zone by 

the flow field induced by pumping and injection wells during ISR or restoration. In general, there 

is a need to know what oxygen levels (and their duration) in the field can be tolerated by the 

biorestoration. This includes determining baseline levels of oxygen and/or nitrate that may limit 

the effectiveness of in situ biorestoration. 

Even under the idealized hydrology and homogeneous solid phase property distribution, 

there are spatially variable distributions of uranium and other components of interest. This is 

because of the interplay between the spatially variable flow field induced by the operation of the 

5-spot well pattern and the reaction kinetics which controls the availability of reactants for 

primary and subsidiary reactions. This results in behaviors that are different from the laboratory 

column experiments even though the same reaction network and similar concentrations are 

employed. These differences include 1) rapid and complete consumption of lactate via the 

sulfate TEAP reaction, and 2) negligible impact of phyllosilicate Fe(III) TEAP reactions (lactateand acetate-based). 

Based on the field-scale biorestoration modeling, injectate water quality can significantly 

affect the TEAP reaction rates. This has implications for the recycling of groundwater from the 

production well as injectate. Assuming the maintenance of anaerobic conditions of the 

produced groundwater, terminal electron acceptors such as sulfate and U(VI) will be 

progressively removed. In particular, the removal of sulfate should reduce the amount of lactate 

needed. Thus, it is possible that the lactate supply can be reduced based on the monitoring of 

the produced groundwater. While this might reduce the potential for biofouling, it will also 

decrease the amount of FeS produced. FeS and other Fe(II) minerals can suppress U(IV) 

reoxidation and remobilization of U(VI) by providing a reservoir of low redox potential 

(Abdelouas et al., 1999). This tradeoff between injectate concentration, biofouling, redox 

potential, and U(VI) bioreduction is a critical design issue for field-scale biorestoration. 

Some of the findings from this study also relate to current ISR practices. 

• It is very important to maintain anoxic conditions in the recirculated water used for

reinjection during restoration. This will facilitate the removal of oxygen from the ore zone 

and the return to low redox potential conditions. 

• The use of pumping and injection wells in an ISR wellfield results in a nonuniform flow 

field. While the zone of influence in the formation for the wellfield certainly has a pore 

volume associated with it, flow rates are location-dependent. Thus, the injection of one 

pore volume of solution into the formation does not mean that the entire matrix pore 

volume is exposed to the injectate. Some parts of the pore volume (e.g., along the most 

direct flow paths between the injection wells and the production well) are exposed to 
7-4 

considerably more injectate than paths with longer travel times. At some locations, 

several pore volumes will have been injected before injectate arrival. 

The uranium bioremediation studies in shallow aquifers, laboratory studies with ISR sediments, 

and hypothetical field biorestoration simulations generally support the potential use of 

engineered biostimulation to catalyze the reduction of uranium and other redox-sensitive metals 

common to post-ISR groundwater. Absent from the weight of evidence for biorestoration of 

uranium ISR sites is a well-conceived field study. The few attempts to stimulate biorestoration 

of post-ISR aquifers via injection of electron donors (e.g., Smith Ranch) have yielded mixed 

results. The anecdotal descriptions of these attempts are insufficient to determine how 

biorestoration was engineered and monitored. We are, however, aware of a recent field 

biorestoration experiment performed at the Kingsville Dome ISR site using dissolved hydrogen 

as the electron donor to stimulate indigenous microorganisms that lowered U(VI) concentrations 

to pre-operational conditions and maintained those concentrations for 2 years (Cabezas et al., 

2011). 

7.1 Considerations

While the results of this study are generally supportive of a biorestoration alternative for 

uranium ISR sites, a significant weakness of the study is the lack of accounting for the natural 

physical and biogeochemical complexities in the subsurface environment. Spatially variable 

property distributions at multiple scales (e.g., intra- and inter-particle, pore- and continuum) lead 

to transport limitations in the delivery of reagents (e.g., lixiviants, reductants, electron donor) as 

well as the leaching of uranium (and other pollutants of interest) from the sediments. 

Superimposed on these physical heterogeneities are complex mineral assemblages and organic 

matter distributions that can have a profound effect on the biogeochemical reactions that 

underlie the bioremediation principle. This is the motivation for the principal recommendation 

from this study: the need for well-conceived field-studies. The recognition here is that the 

unique attributes of each ISR site are likely to prevent a “one size fits all” approach to 

biorestoration. So it is important to develop a sufficiently mechanistic understanding of the sitespecific processes, properties, and conditions controlling biorestoration. This will facilitate 

determinations whether biostimulation can succeed and, if it can, what approach will be the 

most effective for a given site. 

7.2 Potential Approach

The first specific consideration is to routinely measure dissolved oxygen and oxidationreduction potential (ORP) or Eh. This is absolutely critical for the monitoring of redox-sensitive 

processes, yet is not generally performed or included in ISR guidelines for water quality 

indicators (cf., (NRC, 2009a) . These measurements would facilitate the assessment of the 

state of the post-ISR aquifer and potentially restoration effectiveness. Multiparameter sondes 

less than 2 inches in diameter are capable of being deployed at depths up to 250 meters. They 

can include pH, temperature, and specific conductivity measurements, which would provide 

additional information on the biogeochemical response to biostimulation.
7-5 

Another consideration is to conduct laboratory studies using recovered post-ISR sediments 

and groundwater that have preserved the microbial consortia through maintenance of anaerobic 

conditions and refrigeration. The inclusion of laboratory studies was motivated by the 

differences in microbial consortia and behavior observed in the two biostimulated subsurface 

settings: uranium-contaminated groundwater plume in shallow unconfined aquifers and deep 

post-ISR confined aquifer. It is possible that the groundwater dissolved oxygen and metals 

concentrations immediately following the cessation of lixiviant injection may be sufficiently 

elevated to have toxicity effects on the microbial community. Thus, it may be necessary to 

employ a groundwater sweep to lower the initial post-ISR transient of elevated concentrations in

chemical components of concern. In either case, laboratory biostimulation studies will be used 

to identify which electron donor-electron acceptor- microorganism combination is best suited to 

the site-specific situation, as well as potential optimizations in the electron donor delivery (e.g., 

pulsed), concentration, and duration. Possible electron donors include lactate, ethanol, acetate, 

and possibly hydrogen if a carbon source is included. It is very important for the laboratory 

experiments to replicate the biogeochemical conditions (e.g., pH, alkalinity, major ions, 

nutrients, terminal electron acceptors, etc.) as closely as possible. In this study, bicarbonate 

and sulfate concentrations directly impacted uranium behavior and the rate of lactate 

consumption. 

With the current limited understanding of the processes controlling restoration, reactive 

transport modeling can be a useful tool to systematically couple transport and reaction 

processes for the assessment of conceptual process models. Similar to the approach in this 

study, the modeling framework can initially be built based on laboratory column experiments 

targeting the most effective electron donors and associated concentrations. 

The leap to the field scale will require considerable monitoring data to guide the refinement 

of the process and constrain the modeling approach. Monitoring groundwater during 

bioremediation should include: temperature, dissolved oxygen, ORP/Eh, pH, alkalinity, major 

ions, nutrients,and terminal electron acceptors. This is consistent with the need for 

comprehensive field studies that focus on in situ bioremediation at ISR sites. Given the 

complexity of the bioremediation behaviors observed in the column study presented in this 

report (e.g., metabolic lag, inhibition, terminal electron acceptor pools, depletion, etc.), it is 

difficult to conceive an alternative to modeling. Thus, there is a need to increase the availability 

of reactive transport simulators and familiarize more practitioners with their use. One important 

aspect of modeling is that it provides an organizing principle for systematically addressing 

knowledge and data gaps. In this respect, modeling supports a technical basis for data 

collection activities. 

7.2.1 Field Characterization and Monitoring 

In the first NUREG/CR of this series, “Technical Basis for Assessing Uranium 

Bioremediation Performance (Long et al., 2008),” background on uranium bioremediation, 

principles for collecting data, performing experiments and modeling, and a set of prioritized list 

of site information and performance monitoring parameters were identified (Table 7.1). The 

recommendations were based on DOE-sponsored research programs targeting uraniumcontaminated plumes in shallow alluvial aquifers. Most of those recommendations are also 
7-6 

applicable to ISR sites where biorestoration is being considered. In Figure 7.1, we reproduce 

the summary of activities leading to full-scale field deployment. More recently, the U.S. EPA 

released a draft technical report on post-closure monitoring of uranium ISR sites (EPA, 2011), 

which has since been reviewed by their Science Advisory Board. The motivation for the report 

was to update the environmental protection standards to better address groundwater impacts of 

ISR uranium extraction. A principal focus was the sustainability of aquifer conditions at the end 

of restoration. The monitoring objectives (e.g., detection and compliance monitoring, long-term 

stewardship) also overlap with potential biorestoration implementations. We have adapted 

Table 7.1 to include some of the specific recommendations of the EPA report.
7-7 

Table 7.1 Prioritized Information and Monitoring Parameters for Assessment of Bioremediation 

of U(VI). Adapted from Long et al. (2008). 

Necessary Site Information: Uranium Distribution, Magnitude, Form, and Mobility 

Information area/parameter Desired 

Range*

Comments

Form of uranium and

associated mobility/lability

± 30% of 

estimate

Experiments and sediment extractions to identify 

uranium form and potential for future mobility based 

on labile fraction. Evidence for insoluble uranium 

phase or vadose zone sources of uranium are 

particularly important.

Groundwater flow velocity 

(Darcy flux) and direction

±30% of 

estimate

Seasonal and episodic impact critical

Site hydrogeology: hydraulic 

conductivity, porosity, 

dispersivity, hydrofacies

NA Fundamental to both site and process conceptual 

model

Remediation process 

conceptual model

NA Fundamental to prioritization of monitoring parameters

Particle size characteristics

including pore size distribution 

and surface area

NA Reactive surface area, clays, upscaling lab to field

pH,DO, ORP, specific 

conductivity, and temperature

measured at time of 

groundwater sampling in 

background and treatment 

zone using flow-cell with 

multiparameter probe

DO<0.5, 

ORP<0, 

conductivity 

initial increase, 

pH ~ steady

Values used as overall dynamic indicator of impact of 

bioremediation on subsurface geochemistry. Linkage 

of U(VI) concentrations with parameter change 

evidence for bioremediation process conceptual 

model

Aqueous electron acceptors 

and reduction byproducts in 

background and treatment 

zone: nitrate, nitrite, 

ammonium, Mn(IV/II), sulfate, 

sulfide 

NA Significant concentrations of oxygen and/or other 

electron acceptors above the U TEAP on the redox 

ladder must be addressed by the bioremediation 

strategy and their reduction products monitored. 

Sulfur isotopic analyses may provide supplemental 

information.

Fe(III) mineral abundance NA Fe(III) minerals provide sorption sites for Fe(II) & 

U(VI), terminal electron acceptor for iron-reducing 

bacteria, dissolved Fe(II) source 

Fe(II), sulfide measured in 

field at time of sampling for 

U(VI) (up- gradient, treatment 

zone, and down-gradient)

Increasing 

Fe(II); sulfide 

indicator of 

sulfate 

reduction

Maintaining metal reduction may optimize U(VI) 

removal from groundwater; sulfate reduction may 

enhance long-term immobilization in sulfate-rich 

systems

Electron donor concentration 

in treatment zone

>0 Evidence of delivery and treatment zone distribution; 

consumption calculation based on tracer data

Tracer for electron donor >0 in treatment 

zone

Typically Br is used for conservative tracer, accurate 

indication of donor distribution
7-8 

Alkalinity, pCO2 (measured in 

the field)

NA Indicator of carbonate geochemistry, dissolved 

carbonate/bicarbonate forms strong anionic 

complexes with U(VI) to decrease its adsorption and 

increase its solubility and mobility

Concentration and oxidation 

states of metals (primary: U, 

V, Se, Mo, Mn, Fe, As, Ra, 

secondary: Ba, Cd, Cr, Pb, Ni, 

Ag, Th, Hg)

NA Metal contaminants of interest including their redox 

status, TEAP reaction kinetics, solubility and sorption

Depth discrete U(VI) data 

(upper/mid/lower part of 

contaminated zone)

Regulatory 

Compliance 

Criteria

Decreased effectiveness of treatment in the 

uppermost part of the saturated zone may be 

problematic

Major cations and anions NA Provides additional evidence for dominant 

geochemical aqueous complexation and mineral 

solubility reactions

Impact of treatment process 

on groundwater flow directions 

Dependent on 

background 

flow

Provides assurance that groundwater is not rerouted 

around treatment zone

Desirable Performance Monitoring Parameters

Parameter/Method

Desired 

Range or 

Response* Comments

In situ redox status of U using 

in situ sediment incubators 

(ISIs)

Significant 

U(IV) present

Evidence that precipitation of U(IV) is occurring in situ

obtained via differential U extraction.

Microbiological assessment 

using coupons or in situ

incubators 

Shift to metal 

and/or sulfate 

reduction

Evidence for desired in situ microbial respiration 

obtained from deploying coupons or in situ incubators 

in well bores and periodically measuring microbial 

parameters (see text for additional discussion)

Depth-discrete sediment 

sampling/extraction for U, Fe, 

AVS; V, Se, Mo, Mn, As, Ra 

NA Evidence for conversion of terminal electron 

acceptors

Major dissolved gas 

components in groundwater: 

O2, N2, CO2, H2S, methane; 

CO, N2O

NA Evidence for key TEAPs and microbial metabolism

Time-lapse GPR cross-well or 

electrical measurements

Shift in 

geophysical 

attributes in 

zone of 

electron donor

Indicates two-dimensional distribution of electron 

donor, although impact of other transformations on 

geophysical signatures must be assessed and errors 

associated with tomographic inversion procedures 

can ‘smear’ amendment boundary.
7-9 

Table 7.1 (Continued)

Optional Performance Monitoring Parameters

Parameter/Method

Desired 

Range or 

Response* Comments

Depth-discrete data for 

mandatory geochemical 

parameters

NA Characterizes spatial distribution of fundamental 

biogeochemistry in aquifer

Depth-discrete data for 

desirable monitoring 

parameters

NA Characterizes spatial distribution of desired 

biogeochemical reactions in aquifer

Impact of treatment process 

on hydraulic properties

<15% change Documents possible system clogging of pores

Organic and inorganic carbon 

analyses

NA More accurate documentation of natural organic 

carbon sources carbonate geochemistry

Microbiological assessment 

performed directly on 

sampling of treatment zone 

materials 

Shift to metal 

and/or sulfate 

reduction

Measurements directly on groundwater filtrates or 

sediment cores provide "gold standard" assessment 

of microbial community structure (e.g., PLFA, 16S, 

DNA/RNA chip arrays, or functional chip arrays)

In situ redox status of U by 

direct sampling of in situ 

materials

U(IV)/U(VI) measurements on in situ sediments 

provide "ground truth" for U bioreduction; similarly for 

V, Se, Mn

Time-lapse electrical resistivity 

and self potential tomography

NA Can indicate the 3-D distribution of dominant TEAPs

Time-lapse seismic 

tomography

NA Sensitive to gas evolution and secondary mineral 

precipitation

NA = not applicable

*Based on expected values judge to have an impact on MCL.
7-10

Figure 7.1 Summary of activities leading to full-scale bioremediation deployment in the field. 

Hexagons represent characterization steps. Rectangles represent analysis and design steps

(Long et al., 2008). 

Conceptual Model of Future Uranium Behavior

Baseline Risk Assessment for Unmitigated U Fate

Characterization Prerequisites

• U distribution, form, mobility, compliance

• Hydrologic processes controlling U transport

• Geochemistry controlling U mobility

Uranium Bioremediation Bench-Scale Studies

• Sequence and abundance of electron 

acceptors prior to onset of U reduction

• Microbial community structure / function wrt 

electron donor and terminal electron acceptor

• Bioavailability of terminal electron acceptors 

needed for U bioreduction 

• Processes and kinetics controlling reoxidation 

and remobilization of bioreduced U

• Impact of biogeochemical reaction products 

on U bioreduction and subsequent reoxidation

Uranium Bioremediation Conceptual Model / Strategy

• Identify target microorganisms, TEAPs, electron donor

• Design electron donor concentration and delivery in context of 

site-specific hydrology, geochemistry, and biology

• Use simulation to assess design and long-term performance

Uranium Bioremediation Pilot-Scale Studies

• Pre-biostimulation baseline characterization

• Simulation-based monitoring/sampling scheme

• Assessment of in situ performance 

Full-Scale Field Deployment

• Optimized approach

• Optimized monitoring
7-11

7.2.3 Long-Term Monitoring

Post-restoration stability monitoring begins when regulators determiner that restoration is 

complete and a steady-state has been established. The duration of stability phase monitoring is 

specified in the license application but has historically been as short as 6 months. This period 

has been shown to be insufficient at some well fields monitored for longer periods of time, 

where increasing contaminant concentrations were discovered. More recently, the trend has 

been to increase the monitoring period to a year; however, the actual period of stabilization 

can be several years if low permeability zones that are sequestering contaminants or 

lixiviants become significant sources. With this potential behavior, a longer-term 

performance-based approach that matches the site-specific time scales would seem to be 

appropriate.

8-1 

8 Reference List 

Abdelouas, A., Lutze, W., Gong, W., Nuttall, E.H., Strietelmeier, B.A. and Travis, B.J., 2000. 

Biological reduction of uranium in groundwater and subsurface soil. Science of the Total 

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