User manual | ANAEROBIC BIOREMEDIATION OF HEXAVALENT URANIUM IN GROUNDWATER by

ANAEROBIC BIOREMEDIATION OF HEXAVALENT URANIUM IN
GROUNDWATER
by
Aida Tapia-Rodriguez
_____________________
A Dissertation Submitted to the Faculty of the
DEPARTMENT OF CHEMICAL AND ENVIRONMENTAL ENGINEERING
In Partial Fulfillment of the Requirements
For the Degree of
DOCTOR OF PHILOSOPHY
WITH A MAJOR IN ENVIRONMENTAL ENGINEERING
In the Graduate College
THE UNIVERSITY OF ARIZONA
2011
2
THE UNIVERSITY OF ARIZONA
GRADUATE COLLEGE
As members of the Dissertation Committee, we certify that we have read the dissertation
prepared by Aida Tapia-Rodriguez
entitled Anaerobic Bioremediation of Hexavalent Uranium in Groundwater
and recommend that it be accepted as fulfilling the dissertation requirement for the
Degree of Doctor of Philosophy
____________________________________________________________Date: 08/02/11
James A. Field
____________________________________________________________Date: 08/02/11
Maria Reyes Sierra-Alvarez
____________________________________________________________Date: 08/02/11
James Farrell
____________________________________________________________Date: 08/02/11
Jonathan D. Chorover
Final approval and acceptance of this dissertation is contingent upon the candidate's
submission of the final copies of the dissertation to the Graduate College.
I hereby certify that I have read this dissertation prepared under my direction and
recommend that it be accepted as fulfilling the dissertation requirement.
____________________________________________________________Date: 08/02/11
Dissertation Director: James A. Field
____________________________________________________________Date: 08/02/11
Dissertation Director: Maria Reyes Sierra-Alvarez
3
STATEMENT BY AUTHOR
This dissertation has been submitted in partial fulfillment of requirements for an
advanced degree at the University of Arizona and is deposited in the University Library
to be made available to borrowers under rules of the Library.
Brief quotations from this dissertation are allowable without special permission, provided
that accurate acknowledgment of source is made. Requests for permission for extended
quotation from or reproduction of this manuscript in whole or in part may be granted by
the head of the major department or the Dean of the Graduate College when in his or her
judgment the proposed use of the material is in the interests of scholarship. In all other
instances, however, permission must be obtained from the author.
SIGNED: Aida Tapia-Rodriguez
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ACKNOWLEDGEMENTS
At this time, I finally have the opportunity to heartily thank my advisors, Dr. Jim Field
and Dr. Reyes Sierra, whose undeniable expert guidance and openhearted support
encouraged me from the start for the achievement of each of my objectives, essential for
the completion of my dissertation. Besides the unique reward of having worked in their
outstanding research group, I owe them all the empowering, friendship and counseling
they gave me in every stage.
It is an honor for me to also thank Dr. Antonia Luna, who genuinely dedicated to my
mentoring with great skill from the beginning and was always there. I would also like to
express my gratitude to Dr. Wenjie (Alex) Sun for his presence and contributions in
every step of this process.
This dissertation would not have been possible without the aid, support and guidance of
each of my professors and staff at the Department of Chemical and Environmental
Engineering and in the Department of Soil, Water and Environmental Science. Every
portion of expertise shared, as well as the time and effort invested was deeply
appreciated, and became essential for the integral advance of what I could achieve until
now. I would also like to give special thanks and recognition to Philip Anderson at the
University Spectroscopy and Imaging Facilities (USIF), who conducted key XRD
analysis of some of the samples of this research and made available his support at all
times. As well, I want to express my thankfulness to Paul Lee at the Department of
Chemistry for his professional support at carrying out the XPS analysis for this work.
I am highly indebted to my laboratory assistants Virginia Tordable, Angela Athey and
Edgar Olivas, who strongly supported the experimental research of this dissertation. I am
also thankful in so many ways to my friends, especially Irail Cortinas, Valeria Ochoa,
Qais Banihani, Daniela Carvajal, Chris Swanson, Citlali Garcia, for their valuable help
and irreplaceable friendship, as well as to all my past and current friends and colleagues
in the laboratories who have encouraged me in every aspect during my research.
I want to express my deepest thanks to Francisco Gomez and to my parents Estela del
Carmen and Jose Trinidad for their care, presence and encouragement during all this
time.
Finally, I want to acknowledge the Consejo Nacional de Ciencia y Tecnologia
(CONACyT) for the financial support provided during all these years as well as the
supervision at all stages for the completion of this work.
5
DEDICATION
To our Marianita, always present,
to Rebeca, Juan Pablo and Jose Miguel,
to my parents and sisters,
and to Paco.
6
TABLE OF CONTENTS
LIST OF FIGURES ...........................................................................................................11
LIST OF TABLES .............................................................................................................18
ABSTRACT.......................................................................................................................19
CHAPTER 1: INTRODUCTION ......................................................................................21
1.1. Uranium nature and chemistry ............................................................................21
1.2. Uses of uranium...................................................................................................23
1.3. Environmental issues linked to hexavalent uranium ...........................................24
1.4. Health effects of uranium and EPA guidelines ...................................................26
1.5. Uranium remediation strategies...........................................................................28
1.5.1. Physical-chemical remediation approaches ................................................28
1.5.2. Biological remediation ...............................................................................30
1.6. Abiotic interactions of zero-valent iron with uranium ........................................39
1.7. Conclusions .........................................................................................................41
1.8. References ...........................................................................................................41
CHAPTER 2: OBJECTIVES .............................................................................................58
2.1. Aim ......................................................................................................................58
2.2. Specific objectives ...............................................................................................58
CHAPTER 3: ANAEROBIC BIOREMEDIATION OF HEXAVALENT URANIUM IN
GROUNDWATER BY REDUCTIVE PRECIPITATION WITH METHANOGENIC
GRANULAR SLUDGE.....................................................................................................60
3.1. Abstract ...............................................................................................................60
3.2. Introduction .........................................................................................................61
3.3. Materials and Methods ........................................................................................64
3.3.1. Source of biomass ......................................................................................64
3.3.2. Batch experiments ......................................................................................65
3.3.3. U(VI) analysis ............................................................................................68
3.3.4. Respiking of uranyl-chloride ......................................................................69
7
TABLE OF CONTENTS – Continued
3.3.5. Endogenous methane production ...............................................................69
3.3.6. Reoxidation of uraninite .............................................................................70
3.3.7. X-ray diffraction (XRD) .............................................................................71
3.3.8. X-ray photoelectron spectroscopy (XPS) ...................................................72
3.3.9. Sequential extraction ..................................................................................73
3.4. Results .................................................................................................................74
3.4.1. Intrinsic U(VI) reducing potential of granular anaerobic sludge ...............74
3.4.2. Alternative electron donors in the biological reduction of U(VI) ..............77
3.4.3. Sustainability of U(VI) reduction ...............................................................80
3.4.4. Evidence of uranium U(IV) ........................................................................82
3.4.5. Kinetic dependence of U(VI) reduction on biomass concentration ...........85
3.5. Discussion ...........................................................................................................88
3.5.1. Intrinsic uranium reducing activity of methanogenic sludge .....................88
3.5.2. Evidence of uranium reduction ..................................................................89
3.5.3. Effects of the endogenous substrates..........................................................90
3.5.4. Contribution of the electron donors to the intrinsic activity ......................92
3.5.5. Cell density dependence and its association to the intrinsic activity .........93
3.6. Conclusions .........................................................................................................94
3.7. References ...........................................................................................................94
CHAPTER 4: URANIUM BIOREMEDIATION IN CONTINUOUSLY FED UPFLOW
SAND COLUMNS INOCULATED WITH ANAEROBIC GRANULES .....................101
4.1. Abstract .............................................................................................................101
4.2. Introduction .......................................................................................................102
4.3. Materials and Methods ......................................................................................104
4.3.1. Inoculum, pack material and basal media composition ...........................104
4.3.2. Set-up of continuous experiment ..............................................................105
4.3.3. Batch toxicity experiment ........................................................................108
4.3.4. Mass balance and determination of extractable U phases ........................109
4.3.5. Analytical methods ...................................................................................109
4.3.6. Chemicals .................................................................................................110
8
TABLE OF CONTENTS – Continued
4.4. Results ...............................................................................................................111
4.4.1. Reduction of U(VI) in the bioreactor with endogenous substrates (RI)...111
4.4.2. Reduction of U(VI) in the bioreactor with exogenous substrates (RII) ...113
4.4.3. Performance of reactor fed U(VI) and NO3- simultaneously ...................116
4.4.4. U speciation ..............................................................................................120
4.4.5. Impact of U on NO3- reduction.................................................................122
4.5. Discussion .........................................................................................................123
4.5.1. High U removal rates ...............................................................................123
4.5.2. Intrinsic reductive activity anaerobic granular sludge .............................123
4.5.3. Reductive precipitation as the main mechanism for the immobilization .124
4.5.4. U(VI) reduction in the absence of e-donor ...............................................125
4.5.5. Effect of ethanol over the endogenous rate ..............................................127
4.5.6. Impact of NO3- on U bioremediation .......................................................128
4.6. Conclusions .......................................................................................................129
4.7. References .........................................................................................................130
CHAPTER 5: ENHANCEMENT OF HEXAVALENT URANIUM REDUCTION BY
ZERO VALENT IRON WITH A BACTERIAL ENRICHMENT CULTURE ..............139
5.1. Abstract .............................................................................................................139
5.2. Introduction .......................................................................................................140
5.3. Materials and Methods ......................................................................................143
5.3.1. Basal medium ...........................................................................................143
5.3.2. Source of inoculum ..................................................................................144
5.3.3. Batch experiments ....................................................................................144
5.3.3.1. Enrichment process of U(VI)-reducing/ZVI-oxidizing culture.......145
5.3.3.2. Uranium reoxidation........................................................................146
5.3.3.3. Experiment with different electron donors ......................................146
5.3.3.4. Preincubation experiments ..............................................................147
5.3.3.5. Production of H2 by corrosion of Fe0 ..............................................148
5.3.3.6. H2 consumption in presence of ferric hydroxide and magnetite .....148
5.3.3.7. Sulfate reduction by the enrichment culture....................................149
5.3.4. 16S rRNA gene clone libraries .................................................................150
5.3.5. Magnetite synthesis procedure .................................................................151
9
TABLE OF CONTENTS – Continued
5.3.6. Analytical methods ...................................................................................154
5.3.6.1. Soluble U .........................................................................................154
5.3.6.2. Soluble iron .....................................................................................154
5.3.6.3. Hydrogen .........................................................................................155
5.3.6.4. Sequential Extraction ......................................................................155
5.3.6.5. X-ray photoelectron spectroscopy ...................................................156
5.3.6.6. Other determinations .......................................................................157
5.3.7. Chemicals .................................................................................................157
5.4. Results ...............................................................................................................158
5.4.1. Reduction of U(VI) with enrichment culture ...........................................158
5.4.2. Microbial community composition of the enrichment culture .................160
5.4.3. U(VI) transformation ................................................................................161
5.4.4. Alternative electron donors for U(VI) reduction ......................................169
5.4.5. U(VI) reduction by Fe0 preincubated with the enrichment culture ..........171
5.4.6. H2 production during anoxic corrosion of Fe0..........................................179
5.4.7. Use of H2 as an electron donor to reduce Fe(III) by the enrichment
culture .................................................................................................................180
5.5. Discussion .........................................................................................................184
5.5.1. Microbial enhancement of uranium removal by Fe0 ................................184
5.5.2. Reduction of U(VI) as main mechanism of removal ...............................185
5.5.3. No evidence for the direct reduction of U(VI) by enrichment culture .....186
5.5.4. Impact of enrichment culture on Fe0 ........................................................187
5.5.5. Reactive secondary minerals as reductants of U(VI) ...............................188
5.5.6. Evidence of anoxic corrosion of Fe0 ........................................................189
5.5.7. Biological Fe(III) reduction by cathodic H2 and its implication on U(VI)
reduction .............................................................................................................189
5.5.8. Passivation of iron ....................................................................................191
5.5.9. Bacterial population .................................................................................192
5.6. Conclusions .......................................................................................................194
5.7. References .........................................................................................................194
10
TABLE OF CONTENTS – Continued
CHAPTER 6: TOXICITY OF URANIUM TO MICROBIAL PROCESSES IN
ANAEROBIC SLUDGE .................................................................................................208
6.1. Abstract .............................................................................................................208
6.2. Introduction .......................................................................................................209
6.3. Materials and Methods ......................................................................................212
6.3.1. Biomass sources .......................................................................................212
6.3.2. Basal medium ...........................................................................................213
6.3.3. Batch toxicity bioassays ...........................................................................214
6.3.3.1. Methanogenic toxicity bioassays.....................................................216
6.3.3.2. Denitrification toxicity bioassays ....................................................217
6.3.3.3. Toxicity over uranium-reduction activity........................................217
6.3.4. Analytical techniques ...............................................................................219
6.3.5. Chemicals .................................................................................................221
6.4. Results and discussion .......................................................................................221
6.4.1. Methanogenic toxicity ..............................................................................221
6.4.2. Toxicity to denitrifying microorganisms ..................................................225
6.4.2.1. Denitrification with elemental sulfur ..............................................225
6.4.2.2. Denitrification with acetate and H2 as electron donors ...................228
6.4.3. Inhibition of uranium reduction activity ..................................................231
6.5. Conclusions .......................................................................................................236
6.6. References .........................................................................................................237
CONCLUSIONS..............................................................................................................246
REFERENCES ................................................................................................................252
11
LIST OF FIGURES
FIGURE 1.1. Major environmental transport pathways from uranium mill tailings to man
............................................................................................................................................26
FIGURE 1.2. Mechanistic scheme of the possible microbial interactions with U(VI) .....31
FIGURE 1.3. In-situ bioremediation by stimulation of reducing conditions by addition of
an organic electron donor...................................................................................................38
FIGURE 1.4. Conceptual scheme of a permeable reactive barrier (taken from Powell and
Associates, http://www.powellassociates.com/sciserv/3dflow.html) ................................39
FIGURE 3.1. Time course of the reduction of uranium in the presence of different
sources of sludge biomass: Moderate level of endogenous substrate (A. Eerbeek sludge),
and high level of endogenous substrate (B. Nedalco sludge). Legend: ---◊---, Not
inoculated; ---○---, Heat killed sludge with H2; —▲—, Live sludge (no electron donor
added); —●—, Live sludge with H2 ..................................................................................75
FIGURE 3.2. Comparison of the U(VI) bioreduction achieved in the presence of different
electron donors with Eerbeek sludge. ---◊---, Not inoculated; ---■---, Heat killed sludge
with acetate; ---▲---, Heat killed sludge with ethanol; ---●---, Heat killed sludge with H2;
—○—, Live sludge without electron donor added; —■—, Live sludge with acetate; —
▲—, Live sludge with ethanol; —●—, Live sludge with H2. ..........................................79
FIGURE 3.3. Time course of the reduction achieved with repeated respikings of U(VI) to
Nedalco treatments. ---◊---, Not inoculated; ---●---, Heat-killed sludge with H2; —▲—
Live sludge (no electron donor added); —●— Live sludge with H2 .................................81
FIGURE 3.4. Reoxidation of previously bioreduced U(VI) back to U(VI) after addition of
O2 gas mixture in treatments (indicated by the dashed vertical line). ---◊---, Not
inoculated; ---○---, Heat-killed sludge with H2; ⋅⋅⋅⋅●⋅⋅⋅⋅, Live sludge with H2 and O2
(poisoned); —●—, Live sludge with H2 and O2 (non-poisoned); —○—, Live sludge
with H2 (no O2 applied)......................................................................................................83
12
LIST OF FIGURES – Continued
FIGURE 3.5. XRD pattern of a sample of Eerbeek sludge after reduction of several
respikes of uranium during 109 days of incubation. Marks (*) corresponding to the UO2
pattern (JCPDS-ICDD Card #75-0455) are positioned at 2θ = 28.1878, 32.6616, 46.8630,
and 55.5873 ........................................................................................................................84
FIGURE 3.6. Bioreduction of U(VI) at different of Eerbeek sludge biomass
concentrations. ---♦---, Not inoculated; heat-killed sludge with H2: ---∆---, 0.1675 g VSS
L-1; ---□---, 0.335 g VSS L-1; ---◊---, 0.67 g VSS L-1; ---○---, 1.34 g VSS L-1; live sludge
with H2: —▲—, 0.1675 g VSS L-1; —■—, 0.335 g VSS L-1; —♦—, 0.67 g VSS L-1; —
●—, 1.34 g VSS L-1 ...........................................................................................................86
FIGURE 3.7. Correlation of initial zero order rates of U(VI) bioreduction to the biomass
concentration in the presence or absence of added H2 as electron donor. A. Eerbeek
sludge, B. Nedalco sludge. —▲—, Live sludge with H2; ---■---, Live sludge without
any electron donor added; —♦— Killed sludge with H2 ...................................................87
FIGURE 3.8. Specific endogenous methane production by different sludges incubated
during 30 days without any electron donor. —●—, Nedalco sludge; —▲—, Eerbeek
sludge .................................................................................................................................91
FIGURE 4.1. Schematic of sand packed reactor used for the treatment of U(VI). .........106
FIGURE 4.2. Time course of measured U concentration in the influent () and the
effluent (□) of column RI during various periods of operation (period I, stabilization;
period II, dynamic steady state; period III, lowerin g of U concentration). Insert: Zoom in
of the effluent U concentration (□) in the period when it approached the MCL. The
horizontal dotted line in this panel represents the MCL concentration for U in drinking
water (0.13 µM). ..............................................................................................................112
13
LIST OF FIGURES – Continued
FIGURE 4.3. Time course of measured U concentration in the influent () and the
effluent (□) concentrations of column RII during various periods of operation (period I,
stabilization; period II, dynamic steady state; period III, lowering of U concentration).
Insert: Zoom in of the effluent U concentration (□) in the period when it approached the
MCL. The horizontal dotted line in this panel represents the MCL concentration for U in
drinking water (0.13 µM) .................................................................................................115
FIGURE 4.4. Time course of measured U concentration in the influent () and the
effluent (□) concentrations of column RIII during various periods of operation (period I,
stabilization; period IIa, dynamic steady state; period IIb, introduction of NO3-; period III,
lowering of U concentration) ...........................................................................................117
FIGURE 4.5. Evolution over time of the influent () and the effluent (□) concentrations
of NO3- (Panel A) and NO2- (Panel B) in column RIII, after addition of NO3- ...............118
FIGURE 4.6. Average concentration of NO3- (diagonal line fill) and NO2- (empty fill) of
the NOx- in the influent and the effluent. Period I: d 235-250 (4.760 mM NO3-), when
there was a rapid consumption of NO3- and moderate to high formation of NO2-. Periods
IIa and IIb: d 250-317 (4.760 mM NO3-), and d 317-373 (2.124 mM NO3-), when there
was near complete inhibition of NO3- conversion ...........................................................119
FIGURE 4.7. Panel A: Distribution of total recovered U (g) for each column: bottom
fraction (B, diagonal line fill), middle fraction (M, cross hatch fill) and top fraction (T,
empty fill). Insert: Zoom in of the mass recovered in the middle and top fractions. Panel
B: Percentage of nitric acid extracted (solid fill, estimate of U(IV)), bicarbonate extracted
(vertical line fill, estimate of U(VI) adsorbed), and water soluble U fraction (empty fill)
in each layer of each column ...........................................................................................121
FIGURE 4.8. Zero-order rates of NO3- and NO2- removal obtained in batch assay at
different U(VI) concentrations. Legend: NO3- (-●-), NO2- (-○-) .....................................122
FIGURE 5.1. Results from X-ray diffraction (XRD) of synthetic magnetite, with labels
corresponding to the Fe3O4 pattern (JCPDS-ICDD Card #75-1609)...............................153
14
LIST OF FIGURES – Continued
FIGURE 5.2. Evaluation of representative clones obtained from the enrichment culture
by rarefaction analysis .....................................................................................................160
FIGURE 5.3. Phylogenetic tree for the bacteria identified in the co-culture enrichment
with the universal bacteria PCR primer set 27F and 1492R ............................................163
FIGURE 5.4. Distribution of recovered mass of U in a treatment with 5 feedings of 30
µM U(VI) through sequential extraction with H2O (soluble), HCO3 (adsorbed) and HNO3
(reduced) ..........................................................................................................................162
FIGURE 5.5. Time-course of uranium reduction by U-enrichment culture added with Fe0
as electron donor and subsequent uranium reoxidation using O2 as oxidant. The dotted
horizontal line indicate the total amount of U(VI) added ................................................163
FIGURE 5.6. Results from X-ray diffraction (XRD) of U(IV). (A) XRD of uraninite
UO2(s) standard. (B) XRD pattern of a solid sample from a treatment with Fe0 and
enrichment culture after the complete consumption of eight consecutive feedings of 60
µM of U(VI). Labels in both panels correspond to the UO2 pattern (JCPDS-ICDD Card
#73-1715) .........................................................................................................................164
FIGURE 5.7. X-ray photoelectron spectroscopy (XPS) U4f7/2 and U4f5/2 binding energy
spectra for two replicates of the solids from the enrichment culture with Fe0 respiked 5
times with 30 µM of U(VI) (pink line in panels A and B), as well as U(IV) in the form of
UO2 (blue line), U(VI) in the forms of UO2Cl2 (red line) and UO3 (black line) .............166
FIGURE 5.8. Experiment with two concentrations of uranium. Panel A) 30 µM and B) 60
µM U(VI). Legends: ---◊---, Abiotic, no Fe0; ---□---, Biological, no Fe0; —○—, Abiotic
with Fe0; —●—, Biological with Fe0 ..............................................................................168
15
LIST OF FIGURES – Continued
FIGURE 5.9. Time course of uranium with different electron donors in the presence of
microbial co-culture for 30 days of incubation. Legends: ---
---, No inocula, no electron
donor; — —, Endogenous; ---▲---, Abiotic + Fe(II) (pH 6.48); ---■---, Abiotic + H2; --●---, Abiotic + Ethanol; —▲—, Biological + Fe(II) (pH 6.48); —■—, Biological + H2;
—●—, Biological + Ethanol ............................................................................................170
FIGURE 5.10. Time course for uranium (panel A) and Fe2+ (panel B) in treatments
preincubated with MNS media. Legends: —◊—, Abiotic, no Fe0; —▲— Preincubated
inocula; ---▲---, Non-preincubated inocula; —□—, Abiotically-preincubated Fe0; ---□---,
Non-preincubated Fe0; —■—, Abiotically-preincubated Fe0 + non-preincubated inocula;
—●—, Preincubated Fe0 / inocula; ---●---, Non-preincubated Fe0 + non-preincubated
inocula ..............................................................................................................................172
FIGURE 5.11. Time course for uranium (panel A) and Fe2+ (panel B) in treatments
preincubated with MS media. Legends: —◊—, Abiotic, no Fe0; —▲— Preincubated
inocula; ---▲---, Non-preincubated inocula; —□—, Abiotically-preincubated Fe0; ---□---,
Non-preincubated Fe0; —■—, Abiotically-preincubated Fe0 + non-preincubated inocula;
—●—, Preincubated Fe0 / inocula; ---●---, Non-preincubated Fe0 + non-preincubated
inocula ..............................................................................................................................173
FIGURE 5.12. Time course of H2 consumption by the enrichment co-culture with SO42as sole electron acceptor. Legends: ---
---, Abiotic + H2 (no SO42-); ---●---, Abiotic +
SO42- + H2; —○—, Biological + H2 (no SO42-); —●—, Biological + SO42- + H2; ---■---,
Abiotic + SO42- (no H2); —■—, Biological + SO42- (no H2) ...........................................176
FIGURE 5.13. Time course of SO42- consumption by the enrichment co-culture with H2
as electron donor. Legends: ---
---, Abiotic + H2 (no SO42-); ---●---, Abiotic + SO42- +
H2; —○—, Biological + H2 (no SO42-); —●—, Biological + SO42- + H2; ---■---, Abiotic +
SO42- (no H2); —■—, Biological + SO42- (no H2) ...........................................................177
16
LIST OF FIGURES – Continued
FIGURE 5.14. Time course of H2S production by the enrichment co-culture with SO42- as
sole electron acceptor and H2 as electron donor. Legends: ---
---, Abiotic + H2 (no
22SO4 ); ---●---, Abiotic + SO4 + H2; —○—, Biological + H2 (no SO42-); —●—,
Biological + SO42- + H2; ---■---, Abiotic + SO42- (no H2); —■—, Biological + SO42- (no
H2) ....................................................................................................................................178
FIGURE 5.15. Production of H2 during the abiotic (chemical) and biotic anoxic corrosion
of ZVI (100 µM) in the absence of U(VI). Legend: —■—, Fe0 + inocula; —●—, Fe0
without inocula.................................................................................................................180
FIGURE 5.16. (A) Use of H2 by the co-culture in the presence of magnetite (Fe3O4). (B)
Time course of soluble Fe2+ concentration. Legends: ---♦---, Abiotic + H2; ---▲---,
Abiotic + Fe3O4+ H2; —■—, Biological + H2 (no iron); —●—, Biological + Fe3O4 + H2;
---○---, Biological (no iron, no H2); —○—, Biological + Fe3O4 (no H2); --- ---, Abiotic +
Fe3O4 (no H2) ...................................................................................................................182
FIGURE 5.17. (A) Use of H2 by the co-culture in the presence of ferric hydroxide
(Fe(OH)3). (B) Time course of soluble Fe2+ concentration. Legends: ---♦---, Abiotic +
H2; ---▲---, Abiotic + Fe(OH)3 + H2; —■—, Biological + H2 (no iron); —●—,
Biological + Fe(OH)3 + H2; ---○---, Biological (no iron, no H2); —○—, Biological +
Fe(OH)3 (no H2); --- ---, Abiotic + Fe(OH)3 (no H2) .....................................................183
FIGURE 5.18. Schematic of hypothesis of the role of microorganisms in the reduction of
Fe(III) from corrosion products. (1) Anoxic corrosion of Fe0 with formation of H2 and
magnetite (Fe3O4(s)); (2) biotic reduction of Fe(III) in magnetite (Fe3O4(s)) with H2 and
release of soluble Fe2+; (3) formation of Fe(II) secondary minerals, such as siderite
(FeCO3(s)), pyrite (FeS(s)) or vivianite (Fe3(PO4)2(s)), and (4) reduction of U(VI) to
insoluble U(IV) with (a) de-passivated Fe0 or (b) Fe(II) secondary minerals .................192
17
LIST OF FIGURES – Continued
FIGURE 6.1. Toxic effect of increasing uranium(VI) concentrations over the acetoclastic
methanogenic activity of a mixed microbial culture (Eerbeek sludge). (A) Time course of
CH4 concentration (%). Concentration of U(VI) in mM: (—○—), 0; (—∆—), 0.05; (—
□—), 0.2; (—▲—) 0.4; (—
—), 0.6; (—
—), 1.0; (—●—), 2.1. (B) Methanogenic
activity with respect to the initial concentration of U(VI) ...............................................223
FIGURE 6.2. Effect of uranium(VI) concentration on nitrate reduction (A) and nitrite
reduction (B) by a thiosulfate-adapted mixed culture utilizing S0 as electron donor.
Concentration of U(VI) in mM: (—○—), 0; (—□—), 0.005; (—∆—), 0.02; (—▲—)
0.1; (—●—), 0.2; (—
—), 0.4; (—
—), 0.6 ................................................................226
FIGURE 6.3. Role of initial uranium(VI) concentration on the normalized denitrification
activity (respect to the uninhibited control) of (A) a thiosulfate-adapted inoculum using
S0 as electron donor; (B) an anaerobic mixed culture (Eerbeek sludge) utilizing acetate as
electron donor, and (C) an anaerobic mixed culture (Eerbeek sludge) utilizing H2 as
electron donor. Legends: (—●—), NO3- reducing activity; (—○—), NO2- reducing
activity..............................................................................................................................227
FIGURE 6.4. Effect of uranium(VI) concentration on nitrate reduction by an anaerobic
mixed culture (Eerbeek sludge) utilizing (A) acetate as electron donor, and (B) H2 as
electron donor. Concentration of U(VI) (in mM): (—○—), 0; (—∆—), 0.02; (—▲—)
0.1; (—●—), 0.2; (—
—), 0.4; (—
—), 0.6 ................................................................229
FIGURE 6.5. Effect of increasing uranium(VI) concentrations over the U(VI)-reducing
activity of an anaerobic mixed culture (Eerbeek sludge). Concentration of U(VI) in mM:
(—□—), 0.03; (—▲—) 0.2; (—
—), 0.4; (—
—), 0.8; (—●—), 1.0 ........................232
FIGURE 6.6. Impact of increasing initial uranium(VI) concentrations on the rate of
uranium removal by Eerbeek sludge in assays supplied with H2 (—●—) and acetate
(—○—) ............................................................................................................................233
18
LIST OF TABLES
TABLE 1.1. Properties of uranium isotopes ......................................................................23
TABLE 1.2. Summary of the main U(VI)-reducing species .............................................34
TABLE 3.1. Summary of zero-order rates obtained from different anaerobic granular
biofilm inocula used for the biological reduction of U(VI) ...............................................77
TABLE 3.2. Averaged reduction rates for the different electron donors in the presence of
the same concentration of Eerbeek sludge .........................................................................80
TABLE 4.1. Description of the reactor periods in terms of operational parameters and
composition of the influent ..............................................................................................107
TABLE 5.1. Rates of uranium reduction over 20 transfers of the enrichment culture at 30
µM U(VI) .........................................................................................................................159
TABLE 6.1. Summary of experimental conditions applied in the toxicity assays ..........215
TABLE 6.2. Concentrations of uranium causing 20% (IC20), 50% (IC50) and 80% (IC80)
inhibition of the activity of methanogenic and denitrifying microorganisms present in the
biomass sources tested .....................................................................................................224
19
ABSTRACT
Uranium contamination of groundwater from mining and milling operations is an
environmental concern. Reductive precipitation of soluble and mobile hexavalent
uranium (U(VI)) contamination to insoluble and immobile tetravalent uranium (U(IV))
constitutes the most promising remediation approach for uranium in groundwater.
Previous research has shown that many microorganisms are able to catalyze this reaction
in the presence of suitable electron-donors. The purpose of this work is to explore lowcost, effective alternatives for biologically catalyzed reductive precipitation of U(VI).
Methanogenic granular sludge from anaerobic reactors treating industrial wastewaters
was tested for its ability to support U(VI)-reduction.
Due to their high microbial
diversity, methanogenic granules displayed intrinsic activity towards U(VI)-reduction.
Endogenous substrates from the slow decomposition of sludge biomass provided
electron-equivalents to support efficient U(VI)-reduction without external electrondonors. Continuous columns with methanogenic granules also demonstrated sustained
reduction for one year at high uranium loading rates. One column fed with ethanol, only
enabled a short-term enhancement in the uranium removal efficiency, and no
enhancement over the long term compared to the endogenous column. Nitrate, a common
co-contaminant of uranium, remobilized previously deposited biogenic U(IV). U(VI) also
caused inhibition to denitrification. An enrichment culture (EC) was developed from a
20
zero-valent iron (Fe0)/sand packed-bed bioreactor. During 28 months, the EC enhanced
U(VI)-reduction rates by Fe0 compared with abiotic Fe0 controls. Additional experiments
indicated that the EC prevented the passivation of Fe0 surfaces through the use of
cathodic H2 for the reduction of Fe(III) in passivating corrosion mineral phases (e.g.
magnetite) to Fe2+. This contributed to the formation of secondary minerals more
enriched with Fe(II), which are known to be chemically reactive with U(VI). To
determine the toxicity of U(VI) to different populations present in uranium contaminated
sites, including methanogens, denitrifiers and uranium-reducers, experiments were
carried out with anaerobic mixed cultures at increasing U(VI) concentrations. Significant
inhibition to the presence of U(VI) was observed for methanogens and denitrifiers. On
the other hand uranium-reducing microorganisms were tolerant to high U(VI)
concentrations. The results of this dissertation indicate that direct microbial reduction of
U(VI) and microbially enhanced reduction of U(VI) by Fe0 are promising approaches for
uranium bioremediation.
21
CHAPTER 1
INTRODUCTION
1.1. Uranium nature and chemistry
Uranium is a ubiquitous, weakly radioactive element occurring naturally in the
earth crust at an average concentration of 0.0003% (equivalent to 3 ppm) [1], forming
part of minerals in soil and bedrock, or soluble in natural waters [2]. Natural
concentrations in soils range from 0.3 to 11.7 ppm [1]; however, zones of uranium
deposits will contain higher than average concentrations which can reach up to 0.2%,
approximately 3 orders of magnitude more than the normal background levels in soils [2].
Natural surface and ground water concentrations range from 0.03 to 2.1 and from 0.003
to 2.0 ppb, respectively [1]. The natural mobility of uranium in the environment is
affected by wind, streams and volcanic activity, and it can enter groundwater by leaching
from natural deposits [3].
Uranium occurs from the (II) to the (VI) valence states, of which U(IV) and
U(VI) are the most commonly found in nature [4]. Hexavalent uranium (U(VI)) occurs
as uranyl ion (UO22+). Due to its solubility in water and tendency to form complexes with
ligands present in natural waters, uranyl ion corresponds to the most reactive, unstable
22
and mobile form of uranium [4]. The most common ligands of U(VI) in the environment
are: CO32-, PO43-, and OH- [4, 5]. U(IV) (tetravalent uranium) occurs commonly in the
form of oxides, such as uraninite (UO2). Uranous compounds are characterized for their
limited solubility [3].
In aerated solutions, at pH ≤ 2.5 (such as in the acidification caused by the
oxidation of sulfide minerals), uranium exists as free uranyl ion; at pH conditions higher
than 2.5, these ions may start sorbing onto solids, such as Fe-oxy(hydr)oxides, commonly
present in mill tailings [6, 7]. In natural environments at circumneutral pH, uranyl ion
can form soluble complexes with carbonate and phosphate, being the carbonate complex
of uranyl the most prevalent species (in the form of [UO2(CO3)2]2- or [UO2(CO3)3]4-) [7],
creating nearly neutral or anionic ions, which enhances its mobilization [8]. In soil,
uranyl carbonate or uranyl phosphate species can form ternary surface complexes [9],
including inner- or outer- sphere complexes to Fe-(hydr)oxides [9, 10]. Above pH 10,
cationic hydroxide complexes can be also important (UO2OH+, (UO2)2(OH)22+,
(UO2)3(OH)5+) [4, 11].
Uranium has 3 main radioisotopes, which are U-234 (234U), U-235 (235U), and U238 (238U). Every isotope has different radioactive properties [2, 12, 13]. Table 1.1
shows the properties of each radioisotope.
238
U is the most abundant isotope of uranium
(99.27%), as well as the longest-lived, and its radioactivity is limited. On the other hand,
234
U is the least abundant and least long-lived but the highest specific activity, so it
23
contributes with most of the radioactivity of natural uranium [2]. One of the properties of
radionuclides is half-life, which is the time that it takes for half of the isotope to give off
its radiation and decay into another substance. The most radioactive of them is the least
half-lived [1].
Table 1.1. Properties of uranium isotopes.
Isotope
U238
U235
U234
Natural Abundance (%)
99.3
0.72
0.006
Half-life (years)
4.47 x 109
7.04 x 108
2.46 x 105
Specific activity (Bq/g)
12,455
80,011
2.31 x 108
Source: Bleise et al., 2003 [1].
1.2. Uses of uranium
Uranium became important with the development of applications of nuclear
energy in weapons, and more currently the production of commercial fuel for nuclear
power plants, which require uranium enriched in the
235
U-isotope. These uses are based
on the property of uranium of being fissionable (split of the nucleus and therefore a
release of large amounts of energy in the process), and the 235U is preferred since it is the
most fissionable isotope [13]. This leads to wastes rich in 238U but poor in 235U and with
24
lower radioactivity than natural uranium. Minor uses of uranium consist in ceramic and
ornament manufacturing [2]. Depleted uranium is the residue from the enrichment of
uranium (separation of
235
U), with the same chemotoxicity as natural uranium but 60%
less radioactive than the former. Depleted uranium also has several uses such as armor
piercing ammunition [1].
1.3. Environmental issues linked to hexavalent uranium
Human activities such as mining, uranium mill processing and the phosphate
fertilizer industry contribute to uranium present in the biosphere.
From uranium
processing, mill tailings are one of the major concerning causes to uranium
contamination [7].
In mining activities, pitchblende and uraninite (both reduced primary minerals), as
well as carnotite (oxidized uranium in the form of vanadate, a secondary mineral) are
typically mined, and uranium is extracted from ore with acid solutions at in situ leaching
facilities [3]. Since uranium ore usually contains a very small percentage of uranium
(between 0.1% and 0.2%) there is a large leftover fraction.
The slurry containing
unrecoverable metals, minerals, chemicals, organics and process water is discharged
normally to a final storage area (named Tailings Storage Facility, or TSF) [3]. The wet
25
sludge eventually becomes dried out; the mill tailings, or remaining sands, are rich in the
chemicals and radioactive materials that were not removed (including decay products
from uranium), and containing between 50 and 86% of the radioactivity of the original
ore [13].
The historical inappropriate disposal of mill tailings had contributed to a legacy of
significant contamination of soils, surface and ground water. Frequently, the sludge has
been dumped into large manmade basins, which have been often abandoned.
This
practice has left material free to migrate into the environment either by seepage
(leaching) or by being dispersed as blown dust. Furthermore, some catastrophic events
have occurred, such as breakage of dams of a uranium mill tailings disposal containment
structure [7].
Uranium contamination is often accompanied by nitrate and sulfate contamination
as nitric and sulfuric acids were used to extract uranium [14-18]. Sulfate contamination
is also common since it is formed by the oxidation of primary sulfide minerals from acid
mine drainage [3].
The production and use of phosphate fertilizers contribute importantly to uranium
contamination since phosphate-rich rocks contain usually important quantities of uranium
(20-300 ppm) [13], such as the isomorphic substitution of U4+ for Ca2+ occurring in
primary mineral apatite (Ca5PO4) [3].
26
1.4. Health effects of uranium and EPA guidelines
Since uranium is ubiquitous in drinking water, the public experiences a
continuous exposure to uranium at low concentrations. The estimated average daily
intake is from 1 to 2 µg from food and 1.5 µg from water [2]. Figure 1.1 presents the
diverse routes of exposure to uranium.
Figure 1.1. Major environmental transport pathways from uranium mill tailings to man
[19].
27
The primary toxic effect of uranium is chemical, and it is related to its soluble
form; in body fluids uranium is dissolved as uranyl ion (UO22+), which may react with
biological molecules leading to renal impairment [2, 12, 20]. In the blood, uranyl ion can
readily combine with proteins and nucleotides forming stable complexes, due to their
high affinity to phosphate, carbonate and hydroxyl groups. The accumulation of salts in
the proximal tubules for high or chronic exposures may be critical and may lead to
irreversible impairment of kidney function, with nephritis being the most serious
chemical effect of uranium [2].
transport-dependent
and
This is because uranyl ion inhibits both the Na+-
independent
ATP
utilization
as
well
as
oxidative
phosphorylation occurring in the mitochondria of tubular cells [21]. This effect induces
to cellular necrosis by delaying or blocking the cell division process [2], causing atrophy
in the reabsorption function of the renal tubular epithelium, finally leading to renal
impairment [21]. At relatively moderate doses , the cells can be replaced, but these new
cells are histopathologically different [12]. Chelating compounds may be used to prevent
or reduce kidney damage in acute exposures. Administration of bicarbonate has been
used for acute exposures, which can actually complex the uranium and provide alkalinity
to the blood, promoting the excretion through glomerular filtration [2, 22].
On the other hand, there is little evidence of adverse radiation effects due to
exposure to natural levels of uranium. In fact, the acute intake of food or water containing
normal amounts of uranium will not cause a major tendency to cancer, since in the
original form uranium can only emit alpha radiation particles. However, when there is
28
chronic or historical exposure to uranium, the accumulation in the body and subsequent
decay into other radioactive substances contributes to increased cancer risks. This is
because the decay process of uranium can release beta and gamma radiation, which have
higher radiation hazard [1]. Reproductive effects of the different forms of uranium on
humans are currently unknown [23].
The US-EPA drinking water regulation is 30 ppb (based on a wide range of
human and animal health studies), equivalent to 20 pCi/L. This limit protects from
kidney toxicity according to the results obtained in epidemiological studies [12].
1.5. Uranium remediation strategies
1.5.1. Physical-chemical remediation approaches
Efforts have been made during several decades to treat uranium contamination
with cost effective and technically feasible methods. The first studies on the treatment of
uranium consisted in conventional physical-chemical methods, such as ion exchange,
coagulation, lime softening, activated alumina, and reverse osmosis.
29
The most applied technology for the treatment of uranium-contaminated water is
the conventional anion exchange. Studies made on small full-scale anion exchange plants
(55 m3/day), demonstrated that uranium concentrations were reduced up to 99.9% [24].
Zhang and Clifford [25] attained 95% of removal in uranium removal tests from
groundwater with a strong base anion exchange resin. However, the high selectivity for
uranium in this method poses difficulties for the regeneration of the resin. Moreover, and
similarly than for lime softening and reverse osmosis, drinking water levels are difficult
to achieve at high initial uranium levels with anion exchange [12].
Coagulation is another important method for the removal of uranium. Some
laboratory coagulation tests with pond water containing uranium at concentrations of 83
µg L-1 were carried out by dosing ferric and aluminum sulfate at various pH values [26].
Removal efficiencies of 95% were attained at pH 10 and doses of 1.5 to 4 mg L-1 of
aluminum sulfate. In the same study, lime softening attained for 85-90% of removal,
whereas anion exchange allowed 99% of removal, with capacities of 55 mg mL-1 of resin.
Similarly, the removal of uranium was evaluated in a full-scale lake water treatment plant
[27]; it consisted of aluminium and iron coagulation, followed by rapid gravity filtration.
The aluminum coagulation achieved 87% of removal at pH 6, and filtration helped to
further reduce the concentration until reaching 92% of total removal. Iron coagulation
achieved 77% in the first step. Nevertheless, uranium removal to drinking water levels
with this method is strongly dependent on pH and coagulant dose [28].
30
Nanofiltration is another important method, especially to remove uranium
complexes. In a test with five different nanofiltration membranes treating synthetic
solutions containing bicarbonate and uranium at concentrations of 1 mg L-1, important
anionic carbonate complexes of uranium were removed by 90 to 98% at near-neutral pH
conditions[29]. More recent studies on plate module membrane nanofiltration have
indicated that selective removal of uranium over other cations is feasible at low pressure,
but the performance is dependent on the pH (speciation) and ionic strength [30].
Although conventional treatment of uranium contamination consisting of
physical-chemical methods can attain high removal efficiencies, these methods present
several constraints. These include difficulties in regenerating media, accumulation of
chemical wastes, and difficulties to achieve lower than guideline values, especially when
the initial concentration of uranium is high [13, 31].
1.5.2. Biological remediation
The research on alternative approaches to conventional uranium remediation have
become of utmost importance. Recently, the biological approaches have gained attention.
The major mechanisms of interaction between microorganisms and uranium are
presented in Figure 1.2. Among them, biosorption is a passive method that consists in the
31
binding to dead or living microbial biomass through surface complexation (adsorption),
which can be followed by other processes such as absorption [14, 32].
Bioaccumulation is another metabolism-independent process (in contrast to the
energy-dependent bioaccumulation of many other metals) that results from increased
membrane permeability in the presence of the radionuclide [33]. Bioaccumulation by
Arthrobacter and Bacillus sp., isolated from uranium deposits, was reported to be
successful for high levels of uranium; in this way, these microorganisms may be used as
adsorption media in situ [34].
Figure 1.2. Mechanistic scheme of the possible microbial interactions with U(VI) [14].
32
Biomineralization, the precipitation with metabolically generated ligands (such as
phosphate) [35, 36] has been also studied. One of the studies consisted in the addition of
glycerol 2-phosphate to the contaminated aquifer, then amending with a Citrobacter sp.
containing the phosphatase (PhoN) to release the phosphate from the glycerol 2phosphate [35]. Precipitation by the anion released and accumulation of insoluble
uranium as polycrystalline NaUO2PO4 around the cell wall was noticeable [35]; the
presence of NH4+ accelerated the rate of precipitation as NH4UO2PO4.
Reductive precipitation is the biologically catalyzed conversion of soluble UO22+
to insoluble UO2(s) [33, 37-41]. It is considered the most promising mechanism for
sustainable uranium immobilization in aqueous environments. Equation (1) shows the
corresponding reaction. In contrast to other biologically based mechanisms, this latter
process is not limited by a specific number of surface-binding sites and allows the
formation of a highly pure mineral [39].
UO22+ (aq) + 2e-
UO2(s)
Eh0’ = +0.41 V (1)
Although oxidizing conditions are the most prevalent in tailings disposal sites,
reducing conditions could be created by localized confinement and microbial activity [7,
42]. This is generally favored at near-neutral pH (from 5 to 8.5), where uranyl ion is less
mobile and therefore more bioavailable [4].
33
Recent evidence has attributed the reductive precipitation of U(VI) to certain
members of Fe(III)-reducing bacteria [43, 44], including Shewanella spp. [8, 45-48] and
Geobacter spp. [49-53], as well as the sulfate-reducing bacteria Desulfovibrio spp. [38,
54-57]. Whereas some of these microorganisms are able to gain energy from utilizing
uranium as a terminal electron acceptor for energy metabolism [58-61], there are some
others that do not link it to energy-yielding metabolism [54, 58]. Table 1.2. shows the
main U(VI)-reducing species in the environment.
Some studies have suggested different reduction pathways (either respiration of
U(VI) or a fortuitous reduction) depending on the diversity and nature of reductases. The
capacity of obtaining energy for growth from using U(VI) as the terminal electron
acceptor has been reported only for some of the Fe(III)-reducing microorganisms,
specifically S. putrefaciens, G. metallireducens GS-15 [16, 38, 62] and Shewanella algae
[58], attributed to a multicomponent enzyme system in the membrane that can allow
energy yield (production of ATP) and growth. Conversely, although the presence of c3type cytochromes has been confirmed essential for Desulfovibrio spp. to perform U(VI)
reduction [55, 63], they constitute soluble, periplasmic cytochromes that cannot yield
energy for growth.
34
Table 1.2. Summary of the main U(VI)-reducing species.
Population
Genus
Fe(III)-reducing
Geobacter
Shewanella
Species
metallireducens GS-15
[49, 62]
sulfurreducens
[64, 65]
alga BrY
[66, 67]
oneidensis MR-1*
[48, 62]
putrefaciens strain CN32
Anaeromyxobacter
Pseudomonas
Sulfate-reducing
Desulfovibrio
References
dehalogenans strain 2CP-C
[8]
[61, 68]
putida
[69]
sp. CRB5
[69]
sp.
[69]
desulfuricans
[54]
desulfuricans strain G20
[57, 63, 70]
vulgaris strain Hildenborough
ATCC 29579
[71]
sulfodismutans DSM 3696
[71]
baarsii DSM 2075
[71]
sp.
[60, 72]
Desulfomicrobium
norvergicum DSM 1741
[71]
Desulfotomaculum
reducens**
[59]
Desulfosporosinus
orientis DSM 765
[73]
Fermentative
Clostridium
sp.
[74]
Others
Deinococcus
radiodurans R1
[75]
alcalescens***
[76]
Thermoanaerobacter
ethanolicus
[77]
Thermus
scotoductus
[78]
Pyrobaculum
islandicum
[79, 80]
Veillonella
Thermophilic
Hyperthermophilic
*
Previously Alteromonas putrefaciens, then Shewanella putrefaciens MR-1.
**
Also able to utilize Fe(III) as electron acceptor.
***
Previously Micrococcus lactilyticus.
35
Since Fe(III)-reducing processes provide a higher energy yield, Fe(III)-reducers
generally outcompetes sulfate reducers for substrates in the reduction process [16, 44].
Finneran et al. [50] established the preference for electron acceptors based in both their
energy yield as well as in their bioavailability; from this analysis, it was concluded that
although Fe(III)-reduction is more energetically favorable, the bioavailability of this
electron acceptor is constrained by its low solubility, and in this way, the reduction of
U(VI) becomes favorable. Equally, U(VI)-reduction constitutes a larger energy yield
than sulfate-reduction. Nevertheless, it has been observed that, in terms of the rate of
substrate utilization, uranium reduction in the presence of Desulfovibrio spp. is more
kinetically favorable than that with Shewanella spp. [16], which means that the former
can work at low electron donor levels. Diverse microbial communities may be stimulated
differently depending on the free energies offered by the reaction with different electron
donors [68, 81].
The nano-scale size of the biogenic uraninite has been reported by many authors
[48, 82, 83], and its long-term stability – and its rate of dissolution – will depend on the
redox conditions of the aquifer, as well as on pH; at high pH values the presence of
carbonate can complex and dissolve U(VI). Certain complexing ligands and other solids
can either chelate U(VI) (inhibiting reduction) or U(IV) (inhibiting formation of UO2(s)).
In natural water, where carbonate is present and at near neutral pH conditions, U(VI)
becomes strongly bound to carbonate as (UO2)2CO3(OH)3−, UO2CO3o, UO2(CO3)22−, and
36
UO2(CO3)34−, decreasing its sorption to minerals and clays [84]. These conditions impose
significant constraints to the bioavailability and stability of the removal processes [85].
In the same way, a wide variety of factors can lead to the remobilization of
uranium in the environment. Oxygen is well known to readily oxidize U(IV) [86-88],
which represents a constraint to the immobilization of uranium as uraninite. Some
minerals such as pyrite and siderite can protect reduced uranium by consuming the
oxygen supplied by infiltrating groundwater [89, 90]. Since a very low amount of
oxygen is needed to reoxidize uraninite, and oxygenated water is continuously entering
the groundwater system, a realistic long-term stability of reduced uranium in groundwater
must rely on the equilibrium between reduction and oxidation processes that can maintain
soluble uranium in a concentration below the guideline value [88, 89]. Nitrate is another
important oxidizing agent that is commonly found in uranium-contaminated sites. The
presence of nitrate has been found to inhibit the reduction of U(VI) [87, 91, 92]. This has
been linked to the fact that both biological dissimilatory denitrification [93, 94] and its
intermediates – nitrite, nitrous oxide and nitric oxide [95] – can biologically and
chemically promote the oxidation and remobilization of uraninite, respectively, even
under reducing conditions. For example, Thiobacillus denitrificans has been identified
for its ability of oxidizing U(IV) through anoxic nitrate-respiration [94]. Therefore, one
strategy to reduce the concentration of nitrate would be to stimulate these microbial
communities prior to remediation [81].
Fe(III) and Mn(IV) may also decrease the
stability of UO2(s) and reoxidize it [96]. However, the location of some uraninite in the
37
periplasm could partially protect uranium from reoxidation by these electron acceptors
[97]. The size of initial uraninite particles is small enough to become colloidal and
mobile in solution and not really immobilized [82]. The effects of the presence of
competing electron acceptors may be enhanced under these conditions. Hence, the
conditions at which remediation processes work are critical in terms of ensuring the
stability of immobilized uranium [40]. It has been observed that the stability of biogenic
uraninite depends on the rate at which it is formed. When uraninite nanoparticles are
formed as part of a slow-rate reduction process, they become larger and more stable than
when formed at faster rates [98]. This condition may allow slower reoxidation rates of
biogenic uraninite [98].
Additionally, and in spite of the stability disadvantages, the
quick aggregation of uraninite particles into precipitates, as well as the difficulties in
transport of some electron acceptors to the site where reduction occurs in the sediments
are factors that could prevent the reoxidation of uranium [44, 99]. Organic matter and
sulfide minerals consuming some of the electron acceptors present in the matrix, as well
as ternary carbonate complexes, could also contribute to increase the stability of biogenic
UO2(s) [99, 100].
Among the bioremediation techniques that have been applied for the reduction of
uranium in contaminated sites, in situ bioremediation systems are based on enhancing
microbial reduction by using either external or intrinsic electron donors. Figure 1.3
presents a schematic of these processes. These methods are more flexible than pumpand-treat methods, which require fast desorption from soil and rocks and in certain cases
38
increase the temperature of the groundwater [17]. The dosage of electron donor should
be slow release, since bulk addition could result in excessive stimulation of sulfate
reduction and consequent production of sulfide [16].
Figure 1.3. In-situ bioremediation by stimulation of reducing conditions by addition of
an organic electron donor [16].
The method of permeable reactive barriers (PRBs) is presented in Figure 1.4.
PRBs are one of the most favored options for uranium remediation; packed with
components for the reduction and precipitation of uranium, they allow the free passage of
groundwater streams and selectively react with the contaminant of interest, immobilizing
it from water [101].
39
Figure 1.4. Conceptual scheme of a permeable reactive barrier (taken from Powell and
Associates, http://www.powellassociates.com/sciserv/3dflow.html).
1.6. Abiotic interactions of zero-valent iron with uranium
Zero-valent iron (Fe0) is a commonly used media in permeable reactive barriers
for the removal of U(VI) and other contaminants from groundwater [102, 103]. Under
anoxic conditions, Fe0 may undergo a series of corrosion reactions producing H2 (g) and
Fe2+(aq); this reaction is represented by equation (2). Further corrosion results in the
formation of secondary minerals.
40
Fe0(s) + 2H2O(aq)
Fe2+(aq) + H2(aq) + 2OH-(aq)
Eh0’ = -0.44 V (2)
Laboratory scale studies have demonstrated that Fe0 is able to sustain effective
removal of U(VI) in solutions, with three main mechanistic literature hypotheses:
adsorption onto secondary iron oxyhydroxide corrosion products [104]; co-precipitation
with iron oxides [105, 106], or chemical reductive precipitation by Fe0 [102, 107-110].
Reductive precipitation has been confirmed as one of the dominant mechanism for the
removal over other processes [102, 109].
Some advantages of the use of Fe0 for the remediation of U(VI) include the more
thermodynamically favorable reaction with the UO22+/UO2 redox couple [5]. In addition,
since Fe0 is an insoluble material, it could constitute a suitable, slow-release electron
donor for bioremediation systems. In addition, it is known that Fe0 can convert into
materials that could prevent U(IV) from oxidation – such as FeS (mackinawite) – by
providing further reducing electron equivalents [18].
41
1.7. Conclusions
Uranium is an abundant material that can come into contact with groundwater as a
result of high natural levels in soil and bedrock, as well as by some anthropogenic
activities such as mining and milling that are part of the nuclear fuel cycle, and from the
production of phosphate fertilizers. The presence of uranium in drinking water carriages
a serious health risk for living organisms. Several physical-chemical and biological
methods have been applied for the mitigation of uranium from groundwater, being
reductive precipitation from soluble U(VI) to insoluble U(IV) the most reliable method to
stabilize uranium from groundwater. Many microorganisms in the environment have
been linked to U(VI) reduction via an enzymatic role.
Remediation systems have
benefited from the advantages posed by zero-valent iron (Fe0), a potent slow-release
electron donor used for the abiotic removal of U(VI).
Current alternatives for the
remediation of U(VI) in groundwater need to be optimized in order to attain highly
efficient U(VI) immobilization at lower cost and maintenance.
1.8. References
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surfaces. Environmental Science & Technology, 2002. 36(16): p. 3504-3511.
57
106.
Noubactep, C., A. Schoner, and G. Meinrath, Mechanism of uranium removal
from the aqueous solution by elemental iron. Journal of Hazardous Materials,
2006. 132(2-3): p. 202-12.
107.
Gu, B., L. Liang, M.J. Dickey, X. Yin, and S. Dai, Reductive precipitation of
uranium(VI) by zero-valent iron. Environmental Science & Technology, 1998.
32(21): p. 3366-3373.
108.
Abdelouas, A., W. Lutze, E. Nuttall, and W.L. Gong, Remediation of U(VI)contaminated water using zero-valent iron. Comptes Rendus De L Academie Des
Sciences Serie Ii Fascicule a-Sciences De La Terre Et Des Planetes, 1999. 328(5):
p. 315-319.
109.
Farrell, J., W.D. Bostick, R.J. Jarabek, and J.N. Fiedor, Uranium removal from
ground water using zero valent iron media. Ground Water, 1999. 37(4): p. 618624.
110.
Riba, O., T.B. Scott, K.V. Ragnarsdottir, and G.C. Allen, Reaction mechanism of
uranyl in the presence of zero-valent iron nanoparticles. Geochimica Et
Cosmochimica Acta, 2008. 72(16): p. 4047-4057.
58
CHAPTER 2
OBJECTIVES
2.1. Aim
The general aim of this research was to investigate various approaches for the
biological reductive immobilization of hexavalent uranium (U(VI)) in groundwater that
can potentially be applied as low-cost and low-maintenance bioremediation methods.
2.2. Specific Objectives
1. Evaluate the potential of anaerobic granular sludge from upward-flow anaerobic sludge
blanket (UASB) reactors for the reductive precipitation of U(VI) in the presence of
different organic and inorganic electron donors.
2. Assess the performance of a continuous upflow reactor containing granular anaerobic
sludge for the long-term remediation of U(VI)-contaminated water with and without
exogenous electron donor.
59
3. Determine the impact of NO3- on U(VI) remediation by the continuous reactor
containing granular anaerobic sludge.
4. Investigate the mechanisms by which a microbial enrichment culture enhances the
reductive precipitation of U(VI) with Fe0.
5. Study the potential toxicity of soluble U(VI) on different microbial processes including
U(VI)-reduction, methanogenesis and denitrification.
60
CHAPTER 3
ANAEROBIC BIOREMEDIATION OF HEXAVALENT URANIUM IN
GROUNDWATER BY REDUCTIVE PRECIPITATION WITH METHANOGENIC
GRANULAR SLUDGE
3.1. Abstract
Uranium has been responsible for extensive contamination of groundwater due to
releases form mill tailings and other uranium processing waste.
Past evidence has
confirmed that certain bacteria can enzymatically reduce soluble hexavalent uranium
(U(VI)) to insoluble tetravalent uranium (U(IV)) under anaerobic conditions in the
presence of appropriate electron donors.
This paper focuses on the evaluation of
anaerobic granular sludge as a source of inoculum for the bioremediation of uranium in
water. Batch experiments were performed with several methanogenic anaerobic granular
sludge samples and different electron donors. Abiotic controls consisting of heat-killed
inoculum and non-inoculated treatments confirmed the biological removal process. In
this study, unadapted anaerobic granular sludge immediately reduced U(VI), suggesting
an intrinsic capacity of the sludge to support this process. The high biodiversity of
anaerobic granular sludge most likely accounts for the presence of specific
microorganisms capable of reducing U(VI). Oxidation by O2 was shown to resolubilize
61
the uranium. This observation combined with X-ray diffraction evidence of uraninite
confirmed that the removal during anaerobic treatment was due to reductive precipitation.
The anaerobic reduction activity could be sustained after several respikes of U(VI). The
U(VI) removal was feasible without addition of electron donors, indicating that the decay
of endogenous biomass substrates was contributing electron equivalents to the process.
Addition of electron donors, such as H2 stimulated the removal of U(VI) to varying
degrees. The stimulation was greater in sludge samples with lower endogenous substrate
levels. The present work reveals the potential application of anaerobic granular sludge
for continuous bioremediation schemes to treat uranium-contaminated water.
3.2. Introduction
Interest for uranium remediation has increased due to the growing awareness of
contamination at mining and processing sites. The environmental contamination results
from the perturbation of naturally occurring uranium minerals through mining and
processing in the nuclear fuel cycle as well as during phosphate enrichment. Mill tailings
from mining are one of the major causes of extensive uranium contamination due to the
large volume of tailings and previous lack of regulations for their disposal [1]. Uranium
is sometimes accompanied by nitrate and sulfate contamination due to the use of nitric
and sulfuric acids to extract uranium [2], as well as by the oxidation of primary sulfide
62
minerals from acid mine drainage [1]. The primary public health concern of uranium is
its chemical toxicity leading to kidney diseases [3]. The U.S. Environmental Protection
Agency has set the primary drinking water limit for uranium at 30 µg L-1, which protects
the public from kidney toxicity according to the results obtained in epidemiological
studies [4].
Hexavalent uranium (U(VI)) and tetravalent uranium (U(IV)) are the most
common valence states of uranium in nature [5]. U(VI) predominantly occurs in the
uranyl ion form (UO22+) [6], which is the most reactive and mobile form due to its
solubility in water and tendency to form complexes with ligands present in natural waters
such as carbonates [2]. U(IV) is a highly stable and insoluble form of uranium that
generally occurs as the mineral uraninite (UO2(s)).
The levels of uranium in the
environment are, strongly dependent on pH and redox properties of the subsurface
environment.
High removal efficiencies (>90%) can be attained by conventional treatment of
uranium contamination based on physical-chemical methods [7]; however, these methods
have several constraints, such as difficulties regenerating media, chemical wastes, and
difficulties to achieve lower than guideline values, especially when the initial
concentration of uranium is high [8]. Alternative approaches to uranium remediation
utilizing microorganisms are being considered, including biosorption, bioacummulation
and biomineralization [9]. The most accepted biological mechanism for sustainable
63
uranium immobilization in aqueous environments is reductive precipitation, which results
from the biologically catalyzed conversion of soluble UO22+ to insoluble UO2(s).
Since the first discovery of anaerobic microorganisms capable of reducing U(VI)
in the presence of an electron donor [10], a large variety of microorganisms are now
known to carry out the reductive precipitation of U(VI) [9, 11, 12]. The majority of these
microorganisms do not link U(VI)-reduction to energy-gaining processes [11], although
there are reports of some Fe(III)-reducing and sulfate reducing bacteria that can utilize
U(VI) as a terminal electron acceptor for energy metabolism [13]. Particles of uraninite
have been observed to accumulate in the periplasmic space and as deposits occurring
externally on the outer membrane around cells [14]. Since c-type cytochromes have been
localized in zones of U(IV) accumulation, they are hypothesized as required biochemical
components for U(VI) bioreduction [11].
Reductive biotransformation has been considered as an interesting option for
uranium bioremediation at contaminated sites [15]. Diverse microbial communities may
be stimulated differently depending on the free energies offered by the reaction with
different electron donors [16]. H2 is an example of an effective electron donor for the
enzymatic reduction of U(VI) [13]. Ethanol has been also used with success, significantly
stimulating the rate of U(VI)-reduction compared to acetate [17].
Field tests
demonstrated that U could be remediated to below EPA’s limit in groundwater using
64
ethanol as electron donor. Reduction of U(VI) to U(IV) in sediments was confirmed
[18].
Anaerobic microbial biofilms can potentially be considered as inoculum to
promote U(VI) reduction. Anaerobic granular biomass from upward-flow anaerobic
sludge blanket (UASB) reactors used for the high-rate treatment of agro-industrial
wastewater [19] is of special interest. The anaerobic granules are highly settleable, have
high specific anaerobic activities [19, 20] and possess a high level of biodiversity [21].
The scope of this work is to determine the extent at which anaerobic granular sludge can
perform U(VI) reduction, and investigate its potential application in uranium
bioremediation of contaminated groundwater.
3.3. Materials and methods
3.3.1. Source of biomass
Anaerobic granular biofilms – as the source of inocula – were obtained from full
scale upflow anaerobic sludge blanket (UASB) reactors at different wastewater treatment
plants. The different granular sludges used were Aviko (Steenderen, The Netherlands),
from potato starch processing wastewater (0.115 g volatile suspended solids (VSS) g-1
65
wet wt); Eerbeek (Eerbeek, The Netherlands), from recycled paper wastewater (0.135 g
VSS g-1 wet wt); Nedalco (Bergen op Zoom, The Netherlands), from a sugar beet
distillery effluent (0.0654 g VSS g-1 wet wt), and Mahou (Guadalajara, Spain) from beer
brewery wastewater (0.0813 g VSS g-1 wet wt). The specific acetoclastic methanogenic
activity of the sludges were 429, 240, 334 and 408 mg COD g-1 VSS d-1 for Aviko,
Eerbeek, Nedalco and Mahou, respectively. The hydrogenotrophic methanogenic
activities measured in Eerbeek and Mahou sludges were 252 and 207 mg COD g-1 VSS d1
, respectively. All sludge biofilms were stored anaerobically at 4ºC.
3.3.2. Batch experiments
Batch assays were carried out in 160-mL serum flasks (Wheaton, Millville, NJ,
USA), containing 100 mL of liquid - mineral basal media - and 60 mL of headspace. The
basal media used consisted of 5 mg L-1 NH4HCO3, 2 mg L-1 K2HPO4, 2.5 mg L-1
MgSO4·7H2O, 1 mg L-1 Ca(OH)2, 0.33 mg L-1 yeast extract, and trace elements in
concentration: 0.5 µg L-1 H3BO3, 28.0 µg L-1 FeSO4·7H2O, 1.06 µg L-1 ZnSO4·7H2O,
4.15
µg
L-1
MnSO4·7H2O,
2.0
µg
L-1
(NH4)6Mo7O24·4H2O, 1.75
µg
L-1
AlK(SO4)2·12H2O, 1.13 µg L-1 NiSO4·6H2O, 23.6 µg L-1 CoSO4·7H2O, 1.0 µg L-1
Na2SeO3·5H2O, 5.0 µg L-1 Na2WO4·H2O, 1.57 µg L-1 CuSO4·5H2O, 10.0 µg L-1 EDTA,
2.0 µg L-1 resazurin. This media was subsequently adjusted to a pH value of 7.0, and
then provided with NaHCO3 to a final concentration of 59 mM. Sodium bicarbonate was
66
used to buffer the pH of the media, as well as a complexing ligand for U(VI),
representing the natural complexed state of uranium in the groundwater [22]. U(VI) was
provided in the form of uranyl chloride trihydrate (UO2Cl2·3H2O), obtained from
International Bio-Analytical Industries, Inc. (Boca Raton, FL, USA); 4-mL aliquots from
a 10 mM stock solution were added for a final concentration of 0.4 mM U(VI) in each of
the bottles. In previous studies, this concentration was determined to be inside the range
of non-inhibitory concentrations in terms of biological U(VI)-reducing activity.
The granular sludge described above was thoroughly washed in a sieve three
times with MilliQ water. The sludge was immediately weighed and transferred to the
bottles, which already contained basal media and uranium. The assays were amended
with the granular sludge to obtain a final concentration of 0.67 g VSS L-1.
Several electron donors were added individually as follows: 4.0 mM ethanol and
0.2 mM sodium acetate, providing to stoichiometric excesses of 60-times fold and 2times fold, respectively based on e- equivalents. The bottles were flushed with a gas
mixture of N2/CO2 (80:20), first directed to the surface of the liquid with open bottles for
1 min, and then sealed with butyl rubber stoppers and aluminum crimp caps; after this,
the N2/CO2 gas mixture was applied for 4 minutes, this time inserting an inlet and an
outlet needle at the top to replace all the remaining oxygen inside the bottle headspace
and ensure anaerobic conditions throughout the experimental period. For treatments with
H2 as the electron-donor, application of N2/CO2 flushing was carried out as described
67
above before H2 was applied. H2 was applied as a gas mixture of H2/CO2 (80:20) with an
overpressure of 0.8 atm to sealed bottles by inverting the bottle and direct injection to the
liquid phase, in order to provide a final concentration of 19.2 mmol H2 Lliq-1, equivalent
to a 48-fold excess over the stoichiometric requirement of this electron donor based on eequivalents.
Controls for the assay were prepared in order to correctly verify the biological and
electron-donor contribution to the experiment. They consisted of replicated bottles with
heat-killed inoculum, as well as bottles without electron donor and without inoculum. In
the case of controls with heat-killed inoculum, the corresponding amount of sludge was
added to separate bottles with a 10-mL aliquot of MilliQ water 3 days ahead of setting up
the experiment, weighed to account for the water losses and covered with aluminum foil.
These bottles were autoclaved under the following scheme: an initial sterilization was
performed at 121ºC for 1 h, allowed to cool down for 24 h; after accomplishing this step,
lost water was replaced by weight difference, then autoclaved at the same temperature for
30 min, allowing to cool down again. This last step was repeated the next day. Finally,
on the day of the experiment, they were amended with the corresponding amounts of
media, uranium and electron-donor in order to get 100 mL of liquid.
All experiments were performed in duplicated replicates. They were incubated in
the dark at 30ºC on an orbital shaker at 150 rpm. Additionally, the controls described
above were incubated along with the treatments.
68
To evaluate the effects of sludge concentration gradients, the same batch set up
preparation procedure was followed as described above, except the following
concentrations of sludge were used in the different treatments: 0.168, 0.335, 0.67, and
1.34 g VSS L-1. Abiotic controls (heat-killed inoculum and no inoculum), as well as
controls of live inoculum without added electron donor, were prepared accordingly in the
concentrations indicated for the live treatments.
3.3.3. U(VI) analysis
Liquid samples were taken initially and periodically in subsequent days to
measure changes in the soluble uranium concentration. Samples were pipetted into
Eppendorf TM centrifuge tubes and immediately centrifuged at 10,000 rpm (RCF of
10,621 x g) for 10 min. After this step, the supernatant was separated from the tube and
transferred to a 3% HNO3 solution.
Soluble uranium was measured by using an Inductively Coupled Plasma – Optical
Emission Spectrometry (ICP-OES) system model Optima 2100 DV from Perkin-Elmer
TM (Shelton, CT, USA). The detection limit for U(VI) was 0.010 mg L-1. Since this
technique is based on the electromagnetic radiation emission or absorption by an ion in
solution, and since U(VI) is being consumed through redox transformation to an
69
insoluble specie U(IV), the reduction process was monitored by measuring the intensity
of the remaining soluble uranium at a wavelength of 385.958 nm.
3.3.4. Respiking of uranyl-chloride
After completion of U(VI) consumption in the biologically active treatments, in
some experiments fresh 4-mL aliquots of concentrated uranyl chloride (10 mM) were
added by injection to the assay bottles, to a final concentration of 0.4 mM U(VI). The
bottles were flushed with N2/CO2 (80:20) for 4 minutes to closed bottles (with an inlet
and outlet needle, as described previously), to avoid any possible oxygen contamination
inside the bottles. For treatments with H2, The H2/CO2 gas mixture was finally applied
with the same procedure as during in the initial feeding to replenish the electron-donor.
Samples were taken accordingly before and after applying spikes for measurement.
3.3.5. Endogenous methane production
In order to account for the level of endogenous electron donor present in Nedalco
and Eerbeek sludge, the methane production was measured in each case. This operation
consisted in 160-mL bottles amended with 100 mL of basal mineral media and 0.67 g
VSS L-1 of each type of sludge. A duplicate replicate was made for each of them, and no
70
electron donor was added to the bottles.
The bottles were flushed following the
procedure previously described for the batch experiments in order to ensure anaerobic
conditions in the bottles.
Methane gas composition of the headspace was analyzed during the first 30 days
of incubation by gas chromatography (GC) using a Hewlett-Packard 5890 Series II
system (Agilent Technologies, Palo Alto, CA). It was equipped with a flame ionization
detector and a Restek Stabilwax-DA fused silica capillary column (30 m length × 0.53
mm ID, Restek Corporation, Bellefonte, PA). Helium was used as the carrier gas at a
flow rate of 18 mL min-1 and a split flow of 85 mL min-1. The column was operated at
140ºC. The injector port and detector temperatures were 180 and 250ºC, respectively.
3.3.6. Reoxidation of uraninite
After depletion of uranyl chloride in a batch set of live treatments prepared as
described above, a duplicate of treatments was killed by poisoning by adding 1 mL from
a 50 mg L-1 NaN3 solution along with 0.025 g of HgCl3; other duplicate was left intact
(not poisoned) for comparison. Next, a gas mixture consisting in He/CO2/O2 (60:20:20)
was flushed into the headspace of the open bottles for 1 minute, then 4 minutes to closed
bottles, according to the same flushing procedure used for batch experiments. After this,
an overpressure of 1.28 atm (18 psi) of He/CO2/O2 (60:20:20) mixture was applied by
71
inverting bottles and injecting directly to the liquid. This ensured that the concentration
was over 6.15 mmol O2 Lliq-1.
3.3.7. X-ray diffraction (XRD)
X-ray diffraction (XRD) of sludge samples were evaluated to confirm the
formation of the U(IV)-containing mineral uraninite. The treatment used for this study
was incubated with 0.67 g VSS L-1 of Eerbeek sludge and H2 gas as described previously.
Uranyl chloride (UO2Cl2·3H2O) was added at an initial concentration of 0.4 mM, and was
respiked three times; each spike was added when the previous spike was consumed.
After the incubation, the bottles were opened inside an anaerobic chamber (Coy
Laboratory Products, Inc.) and solids were carefully separated from the liquid media by
decanting the liquid. Solids and remaining liquid media were quickly deposited in a 25mL vial and sealed with butyl rubber septa and aluminum caps, and were then subjected
to drying with ultrapure nitrogen (N2) gas. After completely drying, the sample was
ground to a fine powder.
An X-ray powder diffractometer (Scintag XDS 2000,
Cupertino, CA) was used for the measurement of XRD profiles in the powdered samples.
The following parameters were set: wavelength of 1.5406 Å using Cu-Ka1 radiation;
generator settings of 40 kV and 40 mÅ; slits-emitter: 2 mm, 4 mm; receiver: 0.5 mm, 0.3
mm; continuous scan (2θ) from 10º to 70º. Raw data were reduced to net intensity by the
application of a fast Fourier noise filter, background substraction and Ka2 stripping. The
72
JCPDS-ICDD database (International Centre for Diffraction Data) was used for the
identification of the crystal structures in the sample.
3.3.8. X-ray photoelectron spectroscopy (XPS)
After separation from supernatant, solid samples were prepared inside an
anaerobic glove box (3% H2:97% N2) for X-ray photoelectron spectroscopy (XPS)
analysis. Samples were dried by flushing them with ultrapure nitrogen gas and were
transported in sealed vessels from inside the anaerobic glove box into the XPS instrument
to avoid their contact with air. A thin layer of powdered sample was applied onto the
sticky carbon tape. XPS spectra were obtained using an Ultra 165 (Kratos analytical)
XPS spectrometer equipped with Monochromatic Al K-alpha radiation, 1486.6 eV run at
300 W. Survey spectra were acquired with a pass energy of 160 eV and elemental
spectral regions were acquired with at pass energy of 20 eV. Reference stable U
standards (UO3 and UO2) were analyzed in parallel to collect reference peak positions
and peak shapes for the uranium oxidation states (U(VI) and U(IV)). The potential of
beam damage of samples due to reduction of U(VI) during XPS measurement was
minimized (as evidenced with standards) by keeping run times less than 1 h.
73
3.3.9. Sequential extraction
A replicate of six treatments were prepared under same anaerobic conditions for
Nedalco sludge and H2 as electron donor. In this case, an initial feeding of 0.4 mM of
uranyl chloride was applied to the treatments, with an additional respike after the initial
concentration was consumed. Then 1 g of wet solids from of each of these treatments was
added to an Eppendorf tube. Uranium extraction method was adapted from Phillips et al.
[23]. In order to remove soluble uranium, 1 mL of MilliQ water was added to each of
these Eppendorf tubes, followed by vortex-mixing and incubating under stationary
conditions at room temperature for 24 hours under anaerobic conditions. After this
period, the tubes were centrifuged at 10000 rpm for 30 min to separate the solid pellets
from the liquid. Extraction with water was performed twice to ensure all soluble fractions
were removed. Next, a 1-mL sodium bicarbonate (1 M) was used to extract the adsorbed
U(VI) fraction from the residual pellets in the Eppendorf tubes. The tubes were shaken in
a vortex mixer to resuspend all solids. These tubes were left sitting under anaerobic
conditions overnight, and then centrifuged by 10000 rpm and 20 min to separate the solid
pellet from the liquid. Extraction with bicarbonate was performed twice. Finally, nitric
acid (1 N) was applied to recuperate the reduced uranium fraction in the pellet by
oxidation. 1-mL aliquots were added to the Eppendorf tubes under aerobic conditions,
then vortex-mixed to resuspend the pellet and allow full contact with the solution. This
suspension was left sitting for 4 hours, then centrifuged for 10000 rpm and 20 min,
removing the supernatant. The nitric acid step was repeated up to twelve times to ensure
74
all possible reduced fractions were recovered. The recovered supernatants from all of
these steps were analyzed by ICP-OES to obtain the U(VI) concentrations extracted.
3.4. Results
3.4.1. Intrinsic U(VI) reducing potential of granular anaerobic sludge
Batch experiments were performed to test the U(VI) reducing capacity of
anaerobic methanogenic granular sludge from different sources. The U(VI) reducing
activity was tested with live and heat-killed inocula with and without hydrogen as
electron donor. Figure 3.1 shows two examples of results obtained for this test. The loss
of soluble uranium indicated the conversion of U(VI) occurred readily in full treatments
with live inocula and added H2 gas as electron donor. The reduction of U(VI) occurred in
the anaerobic sludge without any apparent lag phase. Conversion of U(VI) also occurred
in the treatments with live inocula and no added electron donor, but the rate of such
reduction was somewhat lower with one of the sludge inocula (Eerbeek); while mostly
not affected with another sludge inoculum (Nedalco). The occurrence of U(VI) reduction
in the absence of added H2 indicates endogenous substrates in the sludge may have been
providing the needed electron equivalents. No conversion of U(VI) was observed in noninoculated and heat-killed inoculum controls, even in the presence of added H2 gas,
75
indicating the reactions observed with live inoculum were catalyzed biologically. The
rate of removal of U(VI) in the live treatments was for the most part constant indicating
zero order kinetics.
Figure 3.1. Time course of the reduction of uranium in the presence of different sources
of sludge biomass: Moderate level of endogenous substrate (A. Eerbeek sludge), and
high level of endogenous substrate (B. Nedalco sludge). Legend: ---◊---, Not inoculated;
---○---, Heat killed sludge with H2; —▲—, Live sludge (no electron donor added); —
●—, Live sludge with H2.
76
The two experiments shown in Figure 3.1 indicate that the Eerbeek and the
Nedalco sludges from different sources had markedly different U(VI) reducing activities
with endogenous substrates corresponding to moderate and high rates observed,
respectively. However the rates of reduction were similar with added H2 as electron
donor, indicating that Nedalco sludge may have had a higher level of endogenous
substrate.
The rate of U(VI) reduction was measured in several other samples of anaerobic
granular sludge. The results of all the experiments are summarized in Table 3.1 by
showing the average zero order rates in treatments with and without added H2 as electron
donor. The table illustrates that of the four different anaerobic sludges tested, only the
Eerbeek sludge had a much lower activity (71%) with endogenous substrate when
compared with the corresponding treatment with added H2. For the remaining sludge
samples, the endogenous rate was only 5 to 19% lower without an electron donor
supplement compared to the treatment with added H2, indicating that the sludge biomass
in most of the sludge samples was sufficient to supply the required electron donor. In the
presence of added H2, the highest rate was observed with Nedalco sludge, corresponding
to a specific activity of 20.5 mg U(VI) g-1 VSS d-1. However all of the other sludge
inocula had similar specific activities which were only 34 to 14% lower than the specific
activity of the Nedalco sludge.
77
Table 3.1. Summary of zero-order rates obtained from different anaerobic granular
biofilm inocula used for the biological reduction of U(VI).
Rate [µM day-1]†
Sludge
Type of wastewater
Electron
name
treated
donor
Eerbeek*
Paper recycle
Nedalco*
Aviko
Mahou
Distillery
Starch processing
Beer brewery
Average
Standard
Rate Ratio
Deviation
(endog./H2)‡
H2
49.4
+ 4.2
None
14.4
+ 0.5
H2
57.8
+ 1.8
None
49.7
+ 2.4
H2
38.0
+ 1.4
None
35.9
+ 0.6
H2
44.9
+ 0.7
None
36.3
+0.4
0.290
0.860
0.945
0.809
*
experiments with this footnote were conducted twice.
†
volumetric zero order rate, each bottle contained 0.67 g VSS L-1
‡
ratio of rate with no added electron donor (endogenous) in numerator over rate with
added H2 in denominator
3.4.2. Alternative electron donors in the biological reduction of U(VI)
The potential of alternative electron donors on the U(VI) bioreduction by
anaerobic granular sludge was investigated.
An experiment was set up to compare
reduced organic compounds, acetate and ethanol, with H2 as electron donors in
stimulating the activity of U(VI) reduction in Eerbeek sludge. The Eerbeek sludge was
78
selected for the experiment since it had the lowest endogenous activity, and thus was
considered the most likely sludge to respond to electron donor addition.
Figure 3.2 shows the concentration of soluble U(VI) decreases readily in all
treatments in which live biomass is present.
As was observed in the previous
experiments, the sludge biomass was observed to have activity in the endogenous
treatment. The various electron donors applied varied in their ability to increase the
U(VI) reduction rate beyond the endogenous rate. A comparison of the reduction rates is
provided in Table 3.2. The best stimulation of the rate (2.7-fold) was observed with H2 as
electron donor. Ethanol caused a moderate stimulation (of 57%) and acetate had no
significant impact on increasing the rate. Removal of uranium was again observed to be
negligible for the non-inoculated control, as well as for the heat-killed inoculum controls
made for each electron donor, confirming that the observed reactions with the live
inoculum were biological in nature.
79
Figure 3.2. Comparison of the U(VI) bioreduction achieved in the presence of different
electron donors with Eerbeek sludge. ---◊---, Not inoculated; ---■---, Heat killed sludge
with acetate; ---▲---, Heat killed sludge with ethanol; ---●---, Heat killed sludge with H2;
—○—, Live sludge without electron donor added; —■—, Live sludge with acetate; —
▲—, Live sludge with ethanol; —●—, Live sludge with H2.
80
Table 3.2. Averaged reduction rates for the different electron donors in the presence of
the same concentration of Eerbeek sludge.
Zero Order Rate [µM day-1] †
Electron donor
added
†
Average
Standard
Deviation
None
16.9
+ 0.6
Acetate
18.5
+ 1.3
Ethanol
26.6
+ 3.1
Hydrogen
45.9
+ 4.2
volumetric zero order rate, each bottle contained
0.67 g VSS L-1
3.4.3. Sustainability of U(VI) reduction
As part of examining the feasibility of applying anaerobic sludge for the
bioreduction of U(VI), it is important to investigate whether the biological reductive
activity can be maintained over an extended period of time. To this end, an experiment
was set up with several of the sludge inocula to demonstrate sustained removal of U(VI)
respiked into the bottles each time the previous allotment of U(VI) was consumed. The
experiments were conducted with and without addition of H2 as electron donor. The
electron donor (H2) was resupplied after each respiking of U(VI). Abiotic controls form
the initial addition of U(VI) were monitored over the entire course of the experiment to
lack of any significant reaction in the absence of biological activity. Figure 3.3 shows an
example of a respike experiment with the Nedalco sludge.
81
Figure 3.3. Time course of the reduction achieved with repeated respikings of U(VI) to
Nedalco treatments. ---◊---, Not inoculated; ---●---, Heat-killed sludge with H2; —▲—
Live sludge (no electron donor added); —●— Live sludge with H2.
The figure illustrates that the initial high U(VI) reducing activity is maintained for
the four consecutive feedings of U(VI) tested. The H2 addition only slightly stimulated
the activity, and the degree of stimulation was not greatly changed after respiking U(VI)
several times. The abiotic controls remained constant over the course of 103 d, indicating
that the reaction observed were biological. These results demonstrate that live sludge
sustained its U(VI) reduction activity over time under anaerobic conditions, and therefore
a continuous system could potentially be applied for the removal of U(VI). In a similar
82
approach, repeated U(VI) spikes were applied to treatments with Eerbeek sludge in the
presence and absence of H2 (results not showed). As expected, treatments with added H2
had a higher rate compared to the endogenous treatment. However, both treatments
continued to reduce U(VI) throughout the three feedings of U(VI) applied.
3.4.4. Evidence of uranium U(IV)
To confirm that U(VI) removal is due to its reduction to insoluble U(IV) an
experiment was set up to demonstrate that oxidation by exposure to O2 could resolubilize
the uranium. O2 was chosen as the oxidant due to its high standard reduction electron
potential compared to other electron acceptors. Figure 3.4 illustrates that when O2 was
introduced at the end of a reduction experiment, the reductively precipitated uranium was
readily oxidized and became resolubilized to the original level of U(VI) added. The
observation reinforces the hypothesis that U(VI) removal during the anaerobic
incubations is due to reduction.
In order to determine whether the reoxidation is
chemical or biological, some assays with pre-reduced uranium were poisoned with a
mixture of NaN3 and HgCl3. Figure 3.4 illustrates that the rate of re-oxidation was
similar irregardless whether the cells from pre-reduced uranium assays were poisoned or
not, indicating that U(IV)-oxidation was due to an abiotic process.
83
Figure 3.4. Reoxidation of previously bioreduced U(VI) back to U(VI) after addition of
O2 gas mixture in treatments (indicated by the dashed vertical line). ---◊---, Not
inoculated; ---○---, Heat-killed sludge with H2; ⋅⋅⋅⋅●⋅⋅⋅⋅, Live sludge with H2 and O2
(poisoned); —●—, Live sludge with H2 and O2 (non-poisoned);
—○—, Live sludge
with H2 (no O2 applied).
XRD spectra of uranium in the sludge also confirmed the presence of U(IV) by
the identification of uraninite (UO2) crystals. The XRD pattern of a sample of the sludge
solid material after four spikings of 0.4 mM U(VI) and 109 days of incubation with
Eerbeek sludge with H2 is shown in Figure 3.5. By superimposing the measured
diffraction pattern of UO2 (JCPDS-ICDD Card #75-0455), it was observed that the angle
2θ of the reflections for UO2 coincides with the position of the peaks in the prepared
sample.
84
Figure 3.5. XRD pattern of a sample of Eerbeek sludge after reduction of several
respikes of uranium during 109 days of incubation. Marks (*) corresponding to the UO2
pattern (JCPDS-ICDD Card #75-0455) are positioned at 2θ = 28.1878, 32.6616, 46.8630,
and 55.5873.
In addition, XPS spectra of anaerobic granular sludge samples revealed U4f 7/2 and
U4f
5/2
photoelectron peaks with binding energies of 380.2 ± 0.2 and 390.9 ± 0.3
respectively. The U4f spectra for U(VI) standard showed a peak with a binding energy
value of 381.7, which is higher than the binding energy values detected in samples. It has
previously been demonstrated that the spectra peaks resulting from the U4f core levels can
be determined as U(VI) or U(IV) species based on the binding energies exhibited [24,
85
25]. Since the binding energy for UO2 is lower than that for UO3, a shift of the U4f peaks
to the lower energies is correlated with the reduction of U(VI) to a lower oxidation state.
Sequential extraction resulted in the following trends: water extraction allowed
the recovery of the 0.04 ± 0.02 % of U(VI), representing the soluble remaining fraction of
U(VI) in the sample, while bicarbonate extraction recovered the 1.39 ± 0.38 % of U(VI)
(adsorbed hexavalent fraction to the solids of the sludge pellet). On the other hand, nitric
acid extraction resulted in a major recovery of 98.57 ± 0.40 % of U(VI), indicating that
uranium was originally present in the sample in the U(IV) valence state, since both water
and bicarbonate extractions produced much lesser recoveries. This supports the argument
that U(IV) was the dominant uranium species in the sample.
3.4.5. Kinetic dependence of U(VI) reduction on biomass concentration
Since anaerobic sludge biomass is the likely source of the enzymatic activity for
driving the removal of uranium, it should follow that the reducing activity is probably a
function of the concentration of biomass present in the treatments. Along this basis, an
experiment was carried out by testing different sludge concentrations for the Eerbeek and
Nedalco inocula, with moderate and high levels of endogenous substrate, respectively. In
both cases, the activity in both the presence and absence of H2 was compared. Figure 3.6
86
shows the time course of the reduction achieved at each biomass concentration with the
Eerbeek sludge in the presence of H2.
Figure 3.6. Bioreduction of U(VI) at different of Eerbeek sludge biomass
concentrations. ---♦---, Not inoculated; heat-killed sludge with H2: ---∆---, 0.1675 g
VSS L-1; ---□---, 0.335 g VSS L-1; ---◊---, 0.67 g VSS L-1; ---○---, 1.34 g VSS L-1; live
sludge with H2: —▲—, 0.1675 g VSS L-1; —■—, 0.335 g VSS L-1; —♦—, 0.67 g VSS
L-1; —●—, 1.34 g VSS L-1.
The graph shows that as the sludge concentration increase, the removal of U(VI)
is more rapid. The initial zero order rates of U(VI) reduction are summarized in Figure
3.7 as a function of sludge concentration in the presence and absence of added H2. These
87
rates are seen to increase proportionally with the sludge concentration, indicating a
dependence of the rate on the sludge concentration. The trend is consistent with the
sludge acting as a biocatalyst of the reaction. As expected, in the case of Eerbeek sludge
the rates were higher in the presence of the electron donor compared to its absence, even
at high sludge concentrations. On the other hand, there was only a small difference in the
rates with and without added electron donor for the Nedalco sludge.
Figure 3.7. Correlation of initial zero order rates of U(VI) bioreduction to the biomass
concentration in the presence or absence of added H2 as electron donor. A. Eerbeek
sludge, B. Nedalco sludge. —▲—, Live sludge with H2; ---■---, Live sludge without
any electron donor added; —♦— Killed sludge with H2.
88
3.5. Discussion
3.5.1. Intrinsic uranium reducing activity of methanogenic sludge
In the present study, previously unexposed methanogenic biofilms from UASB
reactors readily reduced U(VI).
The removal commenced immediately without any
observable lag phase. The U(VI) reducing activity could be sustained as evidenced by
immediate and rapid removal in experiments when additional spikes of U(VI) were
applied. These observations combined with the relatively high specific activities observed
(with values ranging from 13.6 to 20.5 mg U(VI) gVSS-1 d-1) clearly indicate the
existence of an initial intrinsic U(VI)-reducing capacity in the anaerobic sludge granules.
The U(VI)-reducing activity, observed in the presence of unacclimated anaerobic
granular sludge, could be explained by the high biodiversity of microorganisms
commonly
found
in
methanogenic
sludge.
The
biodiversity
includes
some
microorganisms closely related to known bacteria with the ability to catalyze the
reduction of uranium, as indicated by several reviews [9, 11, 12]. Clone libraries of 16S
rDNA conducted on sludge samples used in this study clearly indicate the presence of
well-known U(VI)-reducing microorganisms such as Desulfovibrio and Clostridium spp.
in Eerbeek sludge [21, 26]. Sequences of 16S rDNA recovered from denature gradient
gel electrophoresis bands also indicated the presence of Clostridium in the Nedalco and
Mahou sludges [27, 28]. The literature data support the observed intrinsic capacity of
89
anaerobic sludge from different sources to reduce U(VI). The presence of Desulfovibrio,
indicates that Eerbeek sludge should have H2-dependent sulfate-reducing activity, which
was confirmed by measuring a specific sulfate-reducing activity of 26.35 + 0.015 mg
SO42- g-1 VSS d-1.
3.5.2. Evidence of uranium reduction
To confirm that the mechanism of U(VI) consumption was due to a redox
reaction, reoxidation experiments were conducted. This consisted of the addition of
oxygen to treatments where the uranium was suspected to have been previously
biologically reduced to U(IV). The reduced uranium could be resolubilized by the
addition of O2 as oxidant. The oxidation treatment enabled a full recovery of the initial
soluble U(VI), suggesting that reductive precipitation was the main mechanism of
uranium loss during biological reduction and not other mechanisms such as adsorption.
Reoxidation with O2 occurred both in the absence or presence of live (active) cells. The
lack of any difference indicates that O2 causes a chemical oxidation of biogenic U(IV).
In previous studies, O2 was also shown to readily oxidize previously reduced uranium
[29].
The evidence that U(IV) was formed was also confirmed by three other methods.
XRD demonstrated the presence of U(IV)-containing uraninite crystals in the sludge.
90
XPS confirmed the uranium was in the U(IV) oxidation state, with values agreeing
closely with those reported previously for reduced U(IV) in uraninite [25]. Likewise
uranium was in the sludge was not extracted by bicarbonate but was extracted by nitric
acid, which is consistent with the presence of U(IV) [23].
3.5.3. Effects of the endogenous substrates
In the present study, considerable removal of U(VI) took place in the absence of
added electron donors. This fact may be linked to the presence of certain substrates in
the sludge biomass, which may possibly have electron-donating roles.
In previous
experiments with sediments, it was observed that removal of uranium occurred without
any added electron donor after a certain incubation period [17], and it has been suggested
that organic matter present in the sediment could be implicated as the electron-donor
[13]. In this study, only a few anaerobic granular sludge samples could be stimulated in a
notable way with exogenous electron donors (e.g. Eerbeek), probably because this sludge
type had the lowest endogenous electron donor. The level of endogenous electron donor
was measured in terms of the methane production over 30 d by Nedalco and Eerbeek
sludges. The results showed that the level of endogenous substrate, calculated from the
chemical oxygen demand (COD) equivalent of the methane readings after 30 d
incubation, was higher in the case of Nedalco (166 mg CH4-COD g VSS-1) than that
observed for Eerbeek (131 mg CH4-COD g VSS-1). The measured endogenous substrates
91
in these sludges corresponds to 11 to 14 meq e- L-1 and only 0.8 meq e- L-1 is in fact
required to reduce the added 0.4 mM U(VI). The reason Eerbeek responded more to the
exogenous electron donors (see next heading) is because initially the hydrolysis of the
endogenous biomass in that sludge type was very slow (Figure 3.8). The initial rate of
methane release calculated from the readings retrieved over the first 10 days of
incubation were 8.9 and 3.4 mg CH4-COD g VSS-1 d-1 for Nedalco and Eerbeek sludge,
respectively. These results suggest that the sludge may contain several different complex
organic substrates and degradation intermediates that can contribute in the reduction
process, and that the levels of endogenous substrates may vary greatly depending on the
origin of the wastewater and the operating conditions of the corresponding UASB
system.
Figure 3.8. Specific endogenous methane production by different sludges incubated
during 30 days without any electron donor. —●—, Nedalco sludge; —▲—, Eerbeek
sludge.
92
3.5.4. Contribution of the electron donors to the intrinsic activity
The presence of an electron donor is an important prerequisite for U(VI) reduction
[11]. Some of the electron-donating compounds screened in this study (H2, ethanol and
acetate) have been found to be effective electron donors in previous U(VI)-bioreduction
studies [30]. In particular, H2 has been demonstrated to be a very efficient electron donor
for U(VI) reduction. Examples of bacteria that readily utilize H2 as an electron donor
include Anaeromyxobacter dehalogenans 2CP-C [13], Shewanella spp. [31], and
Desulfovibrio vulgaris [32]. In the present study with the sludge having the lowest
endogenous substrate level (Eerbeek), H2 markedly stimulated U(VI)-reduction, so that
the reaction was complete in only a few days. Even with the other sludges having higher
endogenous substrate levels, H2 showed a significant stimulatory effect, albeit that the
effect was relatively low.
Compared to H2, the other electron donors were less effective. Ethanol had an
observable contribution to the reduction process in Eerbeek sludge, possibly related to the
formation of H2 during the anaerobic degradation of ethanol. Acetate on the other hand,
provided no observable stimulation to U(VI) reduction beyond the reduction observed
with the endogenous substrate. These findings are consistent with previous studies
indicating a lower performance of acetate in the reduction of U(VI) compared to H2 [13].
Also in studies with ethanol, it has been pointed out that its performance was better than
that of acetate [17].
93
3.5.5. Cell density dependence and its association to the intrinsic activity
Reduction of U(VI) occurs only in the presence of live inocula, which suggest a
strict enzymatic character of the reaction. The lack of activity in abiotic and heat killed
controls indicate that no other mechanisms can account for U(VI) consumption. Possible
alternative mechanisms could have been chemical reduction, sorption to cell material and
precipitation of U(VI) with media components such as that known to occur with media
containing high levels of NH4+ and PO43- [33, 34]. The requirement of the presence of
live inocula for the uranium reduction to occur has been recognized in previous works
[11].
Additionally in this study, when the concentration of inocula was increased, the
reduction rate also increased. This observation confirms the biological nature of the
removal process, since this dependence is only seen in the live treatments and not in the
controls having the same amount of killed sludge. The increased rates can be attributed
to the increase in the enzyme concentration with increasing amounts of active sludge.
Mullen et al. [35] also found that increasing concentrations of S. oneidensis MR-1
resulted in an increased capacity to reduce uranium. Similarly, Spear et al. [36] found
that the lag phases observed in their experiments were inversely proportional to cell
concentration. Likewise, they found that by increasing cell concentration, increasing
reduction rates were obtained.
94
3.6. Conclusions
The immediate, rapid and sustained U(VI)-reduction, observed in this study,
suggests that methanogenic anaerobic granular sludge has an innate capacity to support
biological reduction of U(VI).
This phenomenon is most likely due to the natural
occurrence of U(VI)-reducing microorganisms in the sludge. In addition, the decay of
endogenous substrates in anaerobic sludge provides electron equivalents to support
U(VI)-reduction. Exogenous electron donors such as H2, stimulates U(VI)-reduction to
varying degrees. A sludge sample with low endogenous substrate levels was stimulated
the most with H2 addition. Reoxidation, sequential extraction and spectrometric evidence
indicated that the U(VI) removed was converted to insoluble U(IV), confirming that the
predominant removal mechanism was reductive precipitation. On the overall, this study
demonstrates the potential feasibility of utilizing granular sludge from UASB reactors for
bioremediation of uranium-contaminated groundwater either in ex situ reactors or in
permeable reactive barriers.
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ATSDR, Toxicological Profile for Uranium, 1999, Agency for Toxic Substances
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WHO, Uranium in Drinking-water, 2004, World Health Organization.
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Fredrickson, J.K., J.M. Zachara, D.W. Kennedy, M.C. Duff, Y.A. Gorby, S.M.W.
Li, and K.M. Krupka, Reduction of U(VI) in goethite (alpha-FeOOH) suspensions
by a dissimilatory metal-reducing bacterium. Geochimica Et Cosmochimica Acta,
2000. 64(18): p. 3085-3098.
6.
Langmuir, D., Uranium solution-mineral equilibria at low temperatures with
applications to sedimentary ore deposits. Geochimica Et Cosmochimica Acta,
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7.
Baeza, A., M.R. Fernandez de la Campa, M. Herranz, F. Legarda, C. Miro, and A.
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Merroun, M.L. and S. Selenska-Pobell, Bacterial interactions with uranium: An
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10.
Lovley, D.R., E.J.P. Phillips, Y.A. Gorby, and E.R. Landa, Microbial Reduction
of Uranium. Nature, 1991. 350(6317): p. 413-416.
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Microbiology, 2006. 60: p. 149-166.
12.
Suzuki, Y. and T. Suko, Geomicrobiological factors that control uranium
mobility in the environment: Update on recent advances in the bioremediation of
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13.
Marshall, M.J., A.C. Dohnalkova, D.W. Kennedy, A.E. Plymale, S.H. Thomas,
F.E. Loffler, R.A. Sanford, J.M. Zachara, J.K. Fredrickson, and A.S. Beliaev,
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17.
Luo, W.S., W.M. Wu, T.F. Yan, C.S. Criddle, P.M. Jardine, J.Z. Zhou, and B.H.
Gu, Influence of bicarbonate, sulfate, and electron donors on biological reduction
of uranium and microbial community composition. Applied Microbiology and
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Wu, W.M., J. Carley, J. Luo, M.A. Ginder-Vogel, E. Cardenas, M.B. Leigh, C.C.
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T. Mehlhorn, S. Carroll, W.S. Luo, M.W. Fields, B.H. Gu, D. Watson, K.M.
Kemner, T. Marsh, J. Tiedje, J.Z. Zhou, S. Fendorf, P.K. Kitanidis, P.M. Jardine,
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Lettinga, G., A.F.M. Vanvelsen, S.W. Hobma, W. Dezeeuw, and A. Klapwijk,
USE OF THE UPFLOW SLUDGE BLANKET (USB) REACTOR CONCEPT FOR
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Gonzalez-Gil, G., P.N.L. Lens, A. Van Aelst, H. Van As, A.I. Versprille, and G.
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sludge bed reactor. Applied and Environmental Microbiology, 2001. 67(8): p.
3683-3692.
21.
Fernandez, N., E.E. Diaz, R. Amils, and J.L. Sanz, Analysis of microbial
community during biofilm development in an anaerobic wastewater treatment
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98
22.
Elias, D.A., L.R. Krumholz, D. Wong, P.E. Long, and J.M. Suflita,
Characterization of microbial activities and U reduction in a shallow aquifer
contaminated by uranium mill tailings. Microbial Ecology, 2003. 46(1): p. 83-91.
23.
Phillips, E.J.P., E.R. Landa, and D.R. Lovley, Remediation of Uranium
Contaminated Soils with Bicarbonate Extraction and Microbial U(Vi) Reduction.
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Riba, O., T.B. Scott, K.V. Ragnarsdottir, and G.C. Allen, Reaction mechanism of
uranyl in the presence of zero-valent iron nanoparticles. Geochimica Et
Cosmochimica Acta, 2008. 72(16): p. 4047-4057.
25.
Scott, T.B., G.C. Allen, P.J. Heard, and M.G. Randell, Reduction of U(VI) to
U(IV) on the surface of magnetite. Geochimica Et Cosmochimica Acta, 2005.
69(24): p. 5639-5646.
26.
Roest, K., H. Heilig, H. Smidt, W.M. de Vos, A.J.M. Stams, and A.D.L.
Akkermans, Community analysis of a full-scale anaerobic bioreactor treating
paper mill wastewater. Systematic and Applied Microbiology, 2005. 28(2): p.
175-185.
27.
Diaz, E.E., A.J.A. Stams, R. Amils, and J.L. Sanz, Phenotypic properties and
microbial diversity of methanogenic granules from a full-scale upflow anaerobic
sludge bed reactor treating brewery wastewater. Applied and Environmental
Microbiology, 2006. 72(7): p. 4942-4949.
28.
Worm, P., F.G. Fermoso, P.N.L. Lens, and C.M. Plugge, Decreased activity of a
propionate degrading community in a UASB reactor fed with synthetic medium
99
without molybdenum, tungsten and selenium. Enzyme and Microbial Technology,
2009. 45(2): p. 139-145.
29.
Komlos, J., B. Mishra, A. Lanzirotti, S.C.B. Myneni, and P.R. Jaffe, Real-time
speciation of uranium during active bioremediation and U(IV) reoxidation.
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31.
Liu, C.X., Y.A. Gorby, J.M. Zachara, J.K. Fredrickson, and C.F. Brown,
Reduction kinetics of Fe(III), Co(III), U(VI) Cr(VI) and Tc(VII) in cultures of
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by
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Macaskie, L.E., R.M. Empson, A.K. Cheetham, C.P. Grey, and A.J. Skarnulis,
Uranium bioaccumulation by a Citrobacter sp. as a result of enzymically
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100
35.
Mullen, L., V. Klepac-Ceraj, C. Pharino, K. Czerwinski, and M. Polz, Cell
Density Dependent Reduction Kinetics of Hexavalent Uranium by Shewanella
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Spear, J.R., L.A. Figueroa, and B.D. Honeyman, Modeling the removal of
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101
CHAPTER 4
URANIUM BIOREMEDIATION IN CONTINUOUSLY FED UPFLOW SAND
COLUMNS INOCULATED WITH ANAEROBIC GRANULES
4.1. Abstract
Nuclear weapon development and demand for nuclear energy has resulted in a
legacy of environmental issues at uranium mining and milling sites including the
contamination of groundwater with hexavalent uranium (U(VI)) and nitrate. Reductive
precipitation of soluble U(VI) to insoluble tetravalent uranium (U(IV)) containing
minerals is one of the more promising approaches to uranium remediation. The objective
of this study was to evaluate the long-term performance of methanogenic granules for the
continuous treatment of U(VI). For this purpose, three sand-packed columns inoculated
with anaerobic biofilm (7.5 g volatile suspended solids L-1 reactor) were operated with or
without ethanol and one column was exposed to nitrate co-contamination. The columns
were operated for 373 days and efficiently removed U (24 mg L-1) in excess of 99.8%.
No long-term benefit of ethanol addition was observed, suggesting that endogenous
substrates in the biofilm were sufficient to drive the reduction reactions. Nitrate addition
was found to inhibit U(VI) reduction and cause re-oxidation of some U(IV) deposited in
the column. Evidence was also found for inhibition of heterotrophic denitrification by the
102
presence of U(VI) which could be verified in batch tests. Sequential extraction was used
to confirm that most of the removed uranium could be recovered as U(IV) deposited in
the base of the upflow columns. Taken as a whole, the results indicate that methanogenic
biofilms can be reliably applied in bioreactor technology for sustained U removal from
groundwater. However, a suitable pretreatment of nitrate would be needed to avoid its
interference with the performance of the subsequent U removal step.
4.2. Introduction
Release of uranium (U) at mine tailing sites has resulted in widespread
contamination of adjacent groundwater aquifers [1]. The most impacted sites will require
long-term remediation [2, 3]. U can lead to a series of health problems, including the
impairment of the kidneys after long-term exposures [4]. Nitrate (NO3-) is frequently
found as a co-contaminant with U [5], due to use of nitric acid in U ore processing.
Exposure to NO3- also causes human diseases, including the methaemoglobinaemia in
infants [6]. Considering the health risks associated with U and NO3-, there is a great need
to explore cost-effective treatment of both U and NO3- at contaminated sites [7].
U mainly exists in a soluble hexavalent form (U(VI)) as uranyl ion (UO22+) and an
insoluble tetravalent form (U(IV)) typically as a mineral (e.g. uraninite, UO2). Microbial
103
reduction of soluble U(VI) to insoluble U(IV) has been proposed as one possible
approach for remediating groundwater contaminated with U [8, 9]. This redox process is
catalyzed by various groups of bacteria under anaerobic conditions, especially metal- and
sulfate-reducing bacteria [10, 11]. Until now, in situ bioremediation of U based on the
formation of biogenic U(IV) has been applied in low-flowrate processes through the
amendment of additives (e.g. organic carbon compounds) to enhance microbial reductive
activity in groundwater [12, 13].
Preliminary work has indicated that granular anaerobic sludge from methanogenic
bioreactors treating agroindustrial wastewater contained indigenous microorganisms with
high activity towards U(VI) reduction [14]. Additionally, endogenous substrates in the
sludge has been shown to supply electron donor for the bioreduction of U(VI).
Therefore, the application of anaerobic sludge for the high-rate removal of U(VI) is
potentially an attractive bioremediation alternative.
The scope of this study was to evaluate the feasibility of applying granular
methanogenic sludge for the long-term, high-rate bioreduction of U in continuous
columns, and to assess the enhancement of U(VI) reduction with ethanol as an exogenous
electron donor. The present study also aims to investigate the simultaneous treatment of
NO3- and U(VI).
104
4.3. Materials and methods
4.3.1. Inoculum, pack material and basal media composition
Anaerobic granular sludge used to inoculate the columns was obtained from an
upflow anaerobic sludge blanket (UASB) reactor (Nedalco, Bergen op Zoom, The
Netherlands) treating sugar beet distillery effluent. The volatile suspended solids (VSS)
content of the sludge was 6.54% on a wet weight basis. The sludge has a specific
acetoclastic methanogenic activity of 334 mg chemical oxygen demand (COD) as
methane g-1 VSS d-1. The sludge sample was stored anaerobically at 4˚C. Prior to its
addition to each of the columns, the sludge was washed with Milli-Q water.
The sand (2.73 g mL-1 particle density, 1.86 g mL-1 bulk density) used in the study
had a particle size of approximately 0.1 to 0.5 mm (SakrateTM play sand, Bonsal
American Inc., Charlotte, NC). The particle size distribution of this material was 38.4,
35.7, 7.6, 6.4, 9.9 and 1.8 mass % in fractions of > 420, 297-420, 250-297, 177-250, 74177 and < 74 µm, respectively. The sand had a predominant composition of silica (SiO2)
and microcline (triclinic potassium feldspar, KAlSi3O8), based on X-ray diffraction
(XRD) analysis of the original material. It was thoroughly washed with Milli-Q water,
then with a solution of 10% v/v HCl to leach metallic impurities, and finally rinsed with
Milli-Q water. Afterwards, the sand was dried in an oven at 105˚C for 24 hr.
105
The mineral composition of the standard basal medium was the same used in a
previous study [14] and was prepared using Milli-Q water. U(VI) was provided as uranyl
chloride trihydrate (UO2Cl2·3H2O), from a 10 mM stock solution to a final concentration
as indicated in Table 4.1. The media was adjusted to pH 7.0 and then autoclaved at 121°C
for 20 minutes. After autoclaving, filter-sterilized NaHCO3 was added to a final
concentration of 1 g L-1 as buffer and complexing agent for U(VI).
4.3.2. Set-up of continuous experiment
Three laboratory-scale up-flow through columns (0.272 L), labeled RI, RII and
RIII, were operated in parallel (Figure 4.1) under anaerobic conditions. Each column was
packed with a mixture containing 290.8 g of acid washed sand and 25.8 g of wet
anaerobic sludge, which provided a final concentration of 7.5 g VSS L-1 in each column.
The RI column was fed with a basal medium lacking ethanol, which was a control
column to investigate the role of endogenous substrate of the sludge contributing to
biological U(VI) reduction. The RII column was supplied with medium amended with
ethanol as electron donor at an initial concentration of 200 mM and later at 1 mM after
day 18. The RIII column was operated exactly the same as RII for the first 7 months, and
then was supplied with KNO3 at a concentration of 4.76 mM for the rest of the period.
All columns were operated at 25°C. The operation of the columns was sustained for 373
days, equivalent to 684 empty bed volumes, and was divided into different periods based
106
on the changes in empty bed hydraulic retention time (HRT), stabilization, as well as in
the composition of the medium as shown in the Table 4.1. Liquid samples were taken
from the influent and effluent ports for the immediate measurement of pH. Liquid
samples were analyzed for soluble U, NO3-, NO2-, ethanol, and acetate.
Figure 4.1. Schematic of sand packed reactor used for the treatment of U(VI).
107
Table 4.1. Description of the reactor periods in terms of operational parameters and composition of the influent.
Period
I
Stabilization
II
Dynamic steady
state
III
Lowering of U
concentration
Reactor I
Reactor II
Reactor III
HRT (days) *
1.21 + 0.20 (d 0 - 103)
1.24 + 0.18 (d 0 - 119),
0.67 + 0.05 (d 119 - 125)
1.25 + 0.24 (d 0 - 119),
0.74 + 0.04 (d 119 - 125)
E-donor
None
Ia: Ethanol (200 mM, d 0 - 18)
Ib: Ethanol (1 mM, d 18 - 125)
Ia: Ethanol (200 mM, d 0 - 18)
Ib: Ethanol (1 mM, d 18 - 125)
E-acceptor
U (100 µM)
U (100 µM)
U (100 µM)
HRT (days) *
1.20 + 0.06 (d 103 - 119),
0.63 + 0.03 (d 119 - 179),
0.41 + 0.05 (d 179 - 324)
0.63 + 0.05 (d 125 - 179),
0.40 + 0.04 (d 179 - 324)
0.63 + 0.05 (d 125 - 179),
0.40 + 0.04 (d 179 - 324)
E-donor
None
Ethanol (1 mM)
Ethanol (1 mM)
E-acceptor
U (100 µM)
U (100 µM)
IIa: U only (100 µM, d 125 - 235)
IIb: U (100 µM, d 235 – 324) and
nitrate (4.760 mM, d 235 – 317;
2.124 mM, d 317 – 324)
HRT (days) *
0.48 + 0.09 (d 324 - 373)
0.54 + 0.13 (d 324 - 373)
0.50 + 0.06 (d 324 - 373)
E-donor
None
Ethanol (1 mM)
Ethanol (1 mM)
E-acceptor
U (20 µM)
U (20 µM)
U (20 µM) and
nitrate (2.124 mM)
* Average and standard deviation.
108
4.3.3. Batch toxicity experiment
In order to test the toxicity of U(VI) towards heterotrophic denitrifying bacteria, a
batch experiment was performed in 160-mL serum bottles supplied with 100 mL of liquid
and 60 mL of headspace. The medium used for this experiment was the same basal
mineral medium described for the columns but amended with 6.0 mM NO3- and 31.2 mM
ethanol. The treatments were inoculated with mixture sand/sludge from RII, which was
equivalent to 0.32 g VSS L-1. The liquid content of the assay bottles were flushed during
1 min with a mixture of N2/CO2 (80:20, v/v) in order to create anaerobic condition, and
then sealed with butyl rubber septa stoppers and crimp aluminum caps. The closed bottles
were again vigorously flushed during 5 min with the same gas mixture through needles
inserted in the septa to allow the continuous passage of gas through the headspace. The
experiment included abiotic controls incubated with NO3- but without the sand/sludge
mixture. All the treatments were carried out in duplicate and incubated in a static
incubator at 30°C. Liquid samples were taken to determine the NO3- and NO2concentration during the experiment. After complete consumption of the initial amount
of NO3-, a spike of the same concentration of NO3- and ethanol was injected, along with a
range of U(VI) concentrations (0, 100, 400 and 600 µM) provided from a 10 mM U(VI)
stock solution. The concentrations of NO3-, NO2- and U(VI) were monitored.
109
4.3.4. Mass balance and determination of extractable U phases
At the end of the column experiments, the packing materials of each column were
divided into 3 parts (bottom, middle and top) along the vertical profile. Each part was
immediately transferred into capped bottles and flushed with N2 gas to avoid possible reoxidation of U(IV) by air. Solid samples of each bottle were individually homogenized
inside an anaerobic chamber (COY Laboratory Products Inc., Grass Lake, MI).
Subsequently, samples from each layer of each column were taken to determine their dry
weight content and use for sequential U extraction. The extraction was performed on 1 g
of the wet solids from each layer added to an Eppendorf tube. Successive extractions with
MilliQ water, bicarbonate (1.0 M) and nitric acid (1.0 N) were performed to fractionate U
into water soluble U(VI), adsorbed U(VI) and solid phase U(IV), respectively [14].
4.3.5. Analytical methods
Liquid samples and extracts were centrifuged in centrifuge tubes at 10,621x g for
10 min. Soluble U was diluted into a 3% HNO3 solution and measured by using an
inductively coupled plasma optical emission spectrometer (ICP-OES) system model
Optima 2100 DV (Perkin-Elmer
TM
, Shelton, CT, USA) equipped with an AS-93 Plus
autosampler. The wavelength used was 385.958 nm.
110
Ethanol and acetate were measured by gas chromatography (GC) in an Agilent
Technologies 7890A system (Agilent Technologies, Palo Alto, CA) equipped with a
flame ionization detector (FID) and autosampler with an injection volume of 0.5 µL. The
column was a capillary Restek Stabilwax-DA column (30 m length x 0.53 mm ID, Restek
Corporation, Bellefonte, PA). Helium was used as the carrier gas at a flow rate of 30 mL
min-1. The temperatures for the column, injector port and detector were 70, 180, and
275°C. A ramp from 80°C to 120°C at a rate of 20°C min-1 was adapted in the method to
allow the separation of both ethanol and acetate at different retention times.
NO3- and NO2- were measured by suppressed conductivity ion chromatography
(IC) using an IC-3000 system fitted with a Dionex IonPac AS18 analytical column (4 mm
x 250 mm) and an AG18 guard column (4 mm x 40 mm) (Dionex, Sunnydale, CA).
During each injection the eluent (20 mM KOH) was used for 20 min.
Other analytical determinations (e.g., pH, TSS, VSS, etc.) were conducted
according to Standard Methods [15].
4.3.6. Chemicals
Uranyl chloride trihydrate (UO2Cl2·3H2O), obtained from International BioAnalytical Industries Inc. (Boca Raton, FL). Ammonium bicarbonate (NH4HCO3, 21.30-
111
21.73% as NH4+), sodium hydroxide (NaOH), and nitric acid (HNO3, 70%) were all
obtained from Fisher Chemical (Fair Lawn, NJ). Magnesium sulfate (MgSO4·7H2O),
calcium hydroxide (Ca(OH)2), potassium phosphate dibasic (K2HPO4, >99.0%) and
potassium nitrate (KNO3, 99.0%), were all obtained from J.T. Baker (Phillipsburg, NJ).
Yeast extract was purchased from BD (Sparks, MD). Sodium bicarbonate was obtained
from Pfaltz & Bauer (Waterbury, CT). Ethanol (99.5%) and formic acid (HCOOH) were
purchased from Sigma-Aldrich Corp. (St. Louis, MO).
4.4. Results
3.4.1. Reduction of U(VI) in the bioreactor with endogenous substrates (RI)
Figure 4.2 presents the time course of the influent and effluent concentrations of
U of the endogenous column (RI). Steady state removal of U was achieved after 35 d
with an initial loading rate of 86 µmol U(VI) Lr-1 d-1. The U removal efficiency of this
column was stable during the period between d 35-103, with a value of 98.3 + 2.7%
(Figure 4.2). The HRT was decreased from 1.2 to 0.6 d on d 119, and to 0.4 d on d 179
which increased the loading rate to 170 and 246 µmol U(VI) Lr-1 d-1, respectively. The
dynamic steady state period in Figure 4.2 refers to the sustained performance of the
reactor during these highly loaded periods as evidenced by U removal efficiencies of 99.8
112
+ 0.2% (d 119-324). From d 230 onwards, the effluent U concentration started to be
consistently at or below the MCL (0.13 µM) as shown in insert of Figure 4.2.
Figure 4.2. Time course of measured U concentration in the influent () and the effluent
(□) of column RI during various periods of operation (period I, stabilization; period II,
dynamic steady state; period III, lowering of U concentration). Insert: Zoom in of the
effluent U concentration (□) in the period when it approached the MCL. The horizontal
dotted line in this panel represents the MCL concentration for U in drinking water (0.13
µM).
In the final period (from d 324 onwards), the influent concentration of U(VI) was
lowered and maintained at 20 µM (corresponding to 55 µmol U(VI) Lr-1 d-1), which was
113
considered more representative of U concentrations in the field. The reactor continued to
perform effectively with U removal efficiencies of 99.5 + 0.2% (d 324-373). The high
removal efficiencies achieved throughout the operation of column RI proved that the
sludge biomass itself was able to supply sufficient endogenous electron donor to support
the reductive process without the need of an exogenous electron donor.
4.4.2. Reduction of U(VI) in the bioreactor with exogenous substrates (RII)
Figure 4.3 depicts the performance of the column that was fed with ethanol as an
exogenous electron-donor (RII). The column was started up with a large excess of
ethanol (200 mM). This had the effect of promoting an initial achievement of a high U
removal efficiency (98.0 + 1.4%) within 2 d at an initial loading rate of 81 µmol U(VI)
Lr-1 d-1. This was distinctly more rapid than that observed for the endogenous column
(RI) operated at a similar loading. However, the ethanol overloading also caused an
incomplete ethanol bioconversion. The resulting accumulation of acetate caused the pH
of the RII effluent to decrease to 5.0. The acetate in the effluent accounted for 10.9 +
4.6% of the total COD added as ethanol within the initial period of high ethanol loading
(d 0-18). Despite the low pH condition, there were no adverse effects observed on the U
removal activity. However, in order to avoid possible inhibitory effects of acetic acid on
the microbial population, an adjustment was made by lowering the ethanol concentration
from 200 to 1 mM on d 18 (this change is defined by the division Ia/Ib in Figure 4.3).
114
This change allowed the ethanol removal efficiency to increase from 19.6 + 4.1% (d 018) to 96.6 + 4.8% (d 18-119), as well as the amount of acetate formed decreased to
values as low as 1.3 + 0.4% of COD added as ethanol. Additionally, this change allowed
the pH to be restored to the circumneutral range. From d 33 onwards the pH was
consistently controlled at a value of 8.0. Nonetheless, the sudden shift to a lower ethanol
concentration caused an observable disruption in the U removal (Figure 4.3). The U
removal efficiency dropped down to a low of 5.5% by d 20. However, approximately one
month later the column recovered from the disruption and had returned to a new steady
state with an aqueous U removal efficiency of 98.4% + 2.0% (d 49-119) at a loading rate
of 83 µmol U(VI) Lr-1 d-1.
On d 119 and d 179, the HRTs were lowered from 1.2 to 0.6 d and from 0.6 to 0.4
d, which corresponded to an increase in the loading rate to 142 and 239 µmol U(VI) Lr-1
d-1, respectively. These increases in loading rate had no significant impacts on the U(VI)
removal efficiencies, which remained high. The removal of U was 99.7 + 0.13 % (d 119179) and 99.7 + 0.4% (d 179-324), respectively. Between d 225-260, the column was
able to maintain the effluent U concentration below the MCL (insert of Figure 4.3);
however, values consistently above the MCL were observed past day 275.
As was done for RI, the U(VI) concentration in the influent of RII was lowered to
20 µM on day 324 (43 µmol U(VI) Lr-1 d-1). A high U removal continued in this final
115
period, corresponding to 99.1 + 0.44% (d 324-373), with effluent values that bordered the
MCL concentration (Figure 4.3).
Figure 4.3. Time course of measured U concentration in the influent () and the effluent
(□) concentrations of column RII during various periods of operation (period I,
stabilization; period II, dynamic steady state; period III, lowering of U concentration).
Insert: Zoom in of the effluent U concentration (□) in the period when it approached the
MCL. The horizontal dotted line in this panel represents the MCL concentration for U in
drinking water (0.13 µM).
116
4.4.3. Performance of reactor fed U(VI) and NO3- simultaneously
The effect of a concurrent presence of NO3- and U(VI) was investigated in the
third column (RIII). Initially, RIII was amended with ethanol as electron donor, using the
same strategy as described for RII. Figure 4.4 demonstrates a similar performance of
RIII to that of RII when the two reactors were operated similarly (d 0-235). Starting on d
235, the operation of RIII changed with the feeding of 4.76 mM NO3-. The goal was to
achieve simultaneous reduction of both NO3- and U(VI). Almost immediately after the
inclusion of NO3- in the feed, the concentration of U(VI) in the effluent increased swiftly
to concentrations that exceeded somewhat the influent U(VI) concentrations. This
behavior continued to the end of the experiment including the final period when the
influent NO3- and U(VI) concentrations were lowered. The excess U(VI) concentration in
the effluent indicates that immobilized U was becoming mobilized. NO3- was reduced to
nitrite (NO2-) during a 2-week period (Figure 4.5), when approximately 42% of the NO3was converted (Figure 4.6). Thereafter, there was no noteworthy conversion of NO3-,
suggesting a possible inhibition of NO3-- reducing bacteria by prolonged U exposure.
Figures 4.5 and 4.6 also indicate that there was some NO2- in the influent, possibly due to
a conversion NO3- during media storage or in influent lines while being pumped into the
column.
117
Figure 4.4. Time course of measured U concentration in the influent () and the effluent
(□) concentrations of column RIII during various periods of operation (period I,
stabilization; period IIa, dynamic steady state; period IIb, introduction of NO3-; period III,
lowering of U concentration).
118
Figure 4.5. Evolution over time of the influent () and the effluent (□) concentrations of
NO3- (Panel A) and NO2- (Panel B) in column RIII, after addition of NO3-.
119
Figure 4.6. Average concentration of NO3- (diagonal line fill) and NO2- (empty fill) of
the NOx- in the influent and the effluent. Period I: d 235-250 (4.760 mM NO3-), when
there was a rapid consumption of NO3- and moderate to high formation of NO2-. Periods
IIa and IIb: d 250-317 (4.760 mM NO3-), and d 317-373 (2.124 mM NO3-), when there
was near complete inhibition of NO3- conversion.
120
4.4.4. U speciation
At the end of the column experiment (d 373), a sequential extraction was
performed on the lower, middle and upper layers of the sand bed of each column to
estimate the different U species present. Figure 4.7A shows the profile of total recovered
U from each layer. The graph indicates that almost all of the U was immobilized at the
base of each column. The U recovered by extraction was 2.358 and 2.096 g U for RI and
RII, respectively. The cumulative mass of U removed during the reactor operation was
3.836 and 3.664 g U for RI and RII, respectively. This represents values of 61.5 and
57.2% of recovery of U by the sequential extraction compared to the cumulative removal
for RI and RII, respectively. The incomplete U recovery is suspected to be due to a very
steep U concentration gradient at the base of the column, combined with an incomplete
homogenization of the lower layer of the column when collecting the composite samples
for extraction. The total recovery of U by extraction from RIII was 0.244 g U, which was
ca. one order of magnitude lower than from RI and RII, confirming the suspected
remobilization of U(IV) by NO3-. Figure 4.7B shows the speciation of U in each layer.
The plot indicates a large majority of the recovered U is accounted for by the U(IV) in
each of the layers. The U(IV) fraction is greater in the bottom sections of each column.
In the upper sections, a slight increase in the percentage of adsorbed U and water-soluble
species is observed. A noteworthy lower percentage of reduced U was observed in RIII,
the column exposed to NO3-, possibly due to the reoxidation of U(IV).
121
Figure 4.7. Panel A: Distribution of total recovered U (g) for each column: bottom
fraction (B, diagonal line fill), middle fraction (M, cross hatch fill) and top fraction (T,
empty fill). Insert: Zoom in of the mass recovered in the middle and top fractions. Panel
B: Percentage of nitric acid extracted (solid fill, estimate of U(IV)), bicarbonate extracted
(vertical line fill, estimate of U(VI) adsorbed), and water soluble U fraction (empty fill)
in each layer of each column.
122
4.4.5. Impact of U on NO3- reduction
A batch experiment evaluated the inhibitory effects of U(VI) on NO3- and NO2reduction. Figure 4.8 summarizes the zero-order reduction rates of NO3- and NO2- when
exposed to different concentrations of U(VI). Partial to severe inhibition of the NO3reduction rate was occurred in the presence of 400 to 600 µM U(VI). A larger impact of
U(VI) was observed on the rates of NO2- reduction. The concentrations of NO2- that
accumulated in the presence of 100 to 600 µM U(VI) were similar (2.28 to 2.79 mM
NO2-). Nonetheless, the subsequent NO2--removal rates decreased more sharply with
increasing U(VI) compared to the impact on NO3--removal.
Figure 4.8. Zero-order rates of NO3- and NO2- removal obtained in batch assay at
different U(VI) concentrations. Legend: NO3- (-●-), NO2- (-○-).
123
4.5. Discussion
4.5.1. High U removal rates
Only few studies have explored continuous-flow biofilm reactors for the treatment
of U(VI). Most of the previous column studies have used pure bacterial cultures of
Desulfovibrio desulfuricans to attain U reduction [16, 17]. The maximum loading rates
reported have ranged from 8.26 to 105 µmol U(VI) Lr-1 d-1 with efficiencies ranging from
99 to 73% U removal [16, 17]. U(VI) was removed 97.7% from groundwater at a
loading of 1.0 µmol U(VI) Lr-1 d-1 in continuous columns packed with uncontaminated
sediments using ethanol as electron donor [18]. In the present study, loadings of up to
246 µmol U(VI) Lr-1 d-1 (32.8 µmol U(VI) g-1 VSSadded d-1) were effectively treated with
efficiencies as high as 99.8%.
4.5.2. Intrinsic reductive activity anaerobic granular sludge
Non-adapted anaerobic sludge was able to use U(VI) immediately with little or no
lag phase. RII and RIII supplied with ethanol achieved high removal efficiencies within
only 2 days. On the other hand, the endogenous column (RI) gradually improved from
the start, reaching a similar high efficiency after 35 d. The high efficiencies achieved
indicate the presence of microorganisms with an intrinsic ability to biotransform U(VI) in
124
UASB anaerobic sludge granules. Microorganisms taxonomically related to known
bacteria capable of reducing U(VI) [8, 9, 19] have been found in granules of anaerobic
sludge grown in brewery [20-22] or related effluents [23-25]. Sulfate reducers are the
most predominant microorganisms found in granular sludge with a known capacity to
reduce U(VI), in particular Desulfovibrio spp. [20, 23]. Also, Clostridium spp. [20-22]
has been reported to reduce U(VI) [26].
A co-culture of Desulfovibrio spp. and
Clostridia-like organisms – both similar to phylotypes found in granular sludge – was
shown to effectively reduce U(VI) using ethanol as electron donor [27].
4.5.3. Reductive precipitation as the main mechanism for the immobilization
The microbial reduction of U(VI) to U(IV) has been widely reported in literature
as the predominant mechanism of U immobilization. There is widespread evidence that
insoluble U(IV) is formed due to the activity of U(VI)-reducing bacteria, either as
mineral uraninite (UO2) [11, 28-30], or in other forms, such as recently discovered
mononuclear U(IV) [31].
In the present study, the extraction protocol performed
indicated that the majority of U in columns was in the reduced form, U(IV),
corresponding to a removal mechanism of reductive precipitation. Only a small fraction
of the U was released by the bicarbonate extraction step, which should remove adsorbed
U(VI) [32, 33].
125
Biofilms of Desulfovibrio desulfuricans G20 used for the bioremediation of U(VI)
in a continuous-flow reactor were shown to contain black precipitates composed of U(IV)
based on analysis with X-ray absorption near edge structure spectroscopy [34]. The
presence of uraninite in anaerobic granular biofilms which reduced U(VI) in batch
experiments was confirmed with X-ray diffraction [14].
4.5.4. U(VI) reduction in the absence of e-donor
In this study, the long-term removal of U(VI) in the reactor without ethanol
addition (RI) can only be rationalized by the use of biomass as an endogenous electron
donor source. Electron-donor was not limiting, since no long-term benefit was observed
when comparing the endogenous control column with the columns receiving the
continuous supply of ethanol. In the literature, there are very few examples of complex
organic matter serving as electron donor to support U(VI) reduction. Recent studies
showed that anaerobic granular biofilms could readily reduce U(VI) without the addition
of exogenous electron donors [14], and that there was a marginal to moderate benefit of
exogenous electron donor, depending on the source of the granules and type of electron
donor. In addition, only a marginal benefit of added acetate on U(VI)-reduction was
observed in a microcosm study with Chesapeake Bay sediments, suggesting that natural
organic matter (NOM) from the sediments was providing endogenous electron donor
[35]. Oleic acid, an example of a model fatty biomolecule present in biomass, was found
126
to function as a slow-release electron donor for U(VI) reduction [36].
The following
stoichiometry is proposed for biomass as electron donor using a common generic formula
of microbial cells of C5H7NO2 [37]:
UO22+ + ⅟₁₀ C5H7NO2 + ⅘ H2O → UO2 + ½ CO2 + ¹⁹⁄₁₀ H+ + ⅟₁₀ NH4+
(1)
Anaerobic granular biofilms undergo decay. Decay of anaerobic granular biofilms
similar to those used in the current study have been reported to supply an initial flux of
endogenous substrate of 3.43 to 8.94 mg COD g-1 VSS d-1 [14], corresponding to an
organic load of 25.7 to 67.1 mg COD Lr-1 d-1 in the columns of this study based on an
initial biomass concentration of 7.5 g VSS Lr-1 supplied. The highest U(VI) load in this
study was of 246 µmol U(VI) Lr-1 d-1, which corresponds to 3.94 mg COD Lr-1 d-1; thus,
the initial rate of endogenous substrate decay was an order of magnitude higher than that
needed to reduce the highest U load used in this study.
During the entire operation of the columns, a cumulative 16.1 mmol U was
removed from RI. The electron equivalence of the removed U only accounts for 8.9% of
the electron equivalence in the 2.04 g VSS initially supplied to the column. The 30-d
biodegradability of mature granular anaerobic sludge from UASB reactors treating
distillery and waste paper effluents was 9.2 to 11.7% of g COD-CH4/g COD sludge [14].
The measured anaerobic biodegradability of sludge from a UASB treating municipal
wastewater measured over 75 to 129 d was 11 to 32% of g COD-CH4/g COD sludge [38].
127
Given the columns in the present study were operated for 373 d, the biodegradability of
the sludge biomass would be expected to have supplied more than enough electron
equivalence to reduce all the U supplied.
4.5.5. Effect of ethanol over the endogenous rate
The addition of ethanol at high concentrations had a short-term benefit during the
initial period (0-18 d), increasing the U(VI) removal rate in a percentage ~200% over the
endogenous value within 2 days. In previous batch experiments, ethanol was found to
have a small stimulatory effect with granular sludge [14]. Moreover, ethanol stimulated
sediment columns treating U(VI) and Tc(VII) compared with control columns with no
ethanol [18]. Ethanol has also been used as electron donor for reductive precipitation of
U(VI) [36, 39, 40]. In the present work, acetate was an intermediate, consistent with
acetogenic degradation of ethanol to acetate and H2 [41-43].
However, over the long term, no benefit in terms of increased U(VI) removal
efficiency could be distinguished when comparing ethanol-fed versus endogenous
columns. These results contrast previous work with sediment columns where ethanol
addition clearly stimulated bioremediation [18]. However, the difference is most likely
due to the inoculation with biofilm biomass in this study, providing a large pool of
endogenous substrate as a long-term sustained source of electron donor.
128
4.5.6. Impact of NO3- on U bioremediation
Soon after the addition of NO3- to RIII column, the U removal decreased over two
weeks until there was no longer any removal evident. After this period, the effluent
U(VI) concentration was consistently higher than the influent U(VI), indicating a release
of U that was previously immobilized. The most likely mechanism of mobilization is the
chemical oxidation of U(IV) by the NO2- formed from NO3- biotransformation, as shown
by previous literature evidence [33, 44], although there is also evidence for the
microbially catalyzed oxidation of biogenic U(IV) with NO3- [45-48].
After two weeks, NO3- biotransformation also came to a halt, indicating a possible
inhibition of NO3- reduction by U(VI). The inhibition was confirmed in batch tests, in
which the decreases in NO3- and NO2- removal rates occurred at U(VI) of 400 µM and
higher. Only a few previous studies have studied microbial inhibition caused by U.
U(VI) at 3 µM inhibited the degradation of glucose and citrate by Pseudomonas
fluorescens, a common subsurface denitrifying bacterium [49]. Ethanol-driven sulfate
reduction by Desulfovibrio was inhibited by 100 µM of U(VI) [43]. Growth inhibition
of Desulfovibrio desulfuricans G20 at concentrations as low as 140 µM of U(VI) have
been reported [50]. Indirect inhibition (by accumulation of NO2-) is also possible. Nitrite
and free nitrous acid accumulation can result in toxicity to bacteria involved in
denitrification [51]. However, in the present work, the inhibition continued even after
there was a significant decrease in NO2- accumulation.
129
A specific NO3- removal step prior to U bioremediation has been suggested by
previous works [46, 52-54]. The sulfur limestone autotrophic denitrification (SLAD)
process was effective at removing NO3- over long periods of time (4 months) in the
presence of 200 µM U(VI) [53].
4.6. Conclusions
Anaerobic granular biofilm from UASB reactors were effective in removing
U(VI) from water and a high removal capacity (up to 246 µmol U(VI) Lr-1 d-1) could be
sustained for over 1 year. These are the highest volumetric rates reported to date in
biofilm reactors. Most of the removed U could be recovered as U(IV) immobilized at the
base of each column. Addition of an exogenous electron donor, ethanol, had only a shortterm beneficial impact. Over the long-term, endogenous substrates released from biomass
decay supplied sufficient electron donor to support the long-term reduction of U(VI).
NO3- disrupted U removal, due to the reoxidation of immobilized U, and exposure of U
was also responsible for inhibition of denitrification. Thus pretreatment of NO3- in cocontaminated sites is recommended.
130
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139
CHAPTER 5
ENHANCEMENT OF HEXAVALENT URANIUM REDUCTION BY ZERO VALENT
IRON WITH A BACTERIAL ENRICHMENT CULTURE
5.1. Abstract
Zero-valent iron (Fe0) has been used as reactive packing in permeable reactive
barriers to remediate uranium contamination in groundwater. The main mechanism of
removal has been attributed to the reductive precipitation of hexavalent uranium (U(VI))
to insoluble tetravalent uranium (U(IV)). The objective of this work is to determine if
microorganisms can enhance the rate of U(VI)-reduction by Fe0. An enrichment culture
(EC) was developed by serial transfer of 30 µM U(VI) and 1 mM Fe0. Additional batch
experiments were set-up to elucidate mechanisms for the enhancement and the
physiological role of the EC. EC inoculated cultures enhanced U(VI)-reduction by Fe0 3to 27.4-fold compared to abiotic incubations. Sequential extraction and XPS confirmed
that the immobilized U was U(IV). The predominate members of the EC based on a 16S
rRNA gene clone library were closely related to Dechloromonas and Stenotrophomonas.
H2 and Fe2+ were the main anoxic corrosion products of Fe0, yet neither were electron
donors that could be used by EC to reduce U(VI) directly. Preincubation of Fe0 with EC
increased reactivity of Fe0 with U(VI) compared to fresh Fe0, whereas preincubation of
140
Fe0 in abiotic conditions caused complete passivation of Fe0. The EC was able to use
cathodic H2 was an electron donor to reduce Fe(III) in secondary minerals such as
magnetite. The results indicate that EC enhances U(VI) reduction by preventing Fe0
passivation and/or generating biogenic secondary minerals that are more reactive with
U(VI) than those formed under abiotic conditions. An important physiological role of the
EC is to reduce Fe(III) in magnetite and other Fe(III) containing minerals.
5.2. Introduction
Widespread uranium contamination of groundwater in the US has resulted from a
legacy of uranium mining and processing [1, 2]. Serious chemical health effects, such as
damage to the kidneys, can result from long-term exposure to soluble uranium [3]. In the
environment, uranium exists in one of two main valence states. Hexavalent uranium
(U(VI)) is the predominant species under oxidizing conditions, which occurs mainly as
the soluble uranyl ion (UO22+). Tetravalent uranium (U(IV)) is stable under reducing
conditions and occurs as a solid mineral, commonly as uraninite (UO2). A promising
uranium remediation approach involves the reduction of U(VI) to insoluble U(IV). Zerovalent iron (Fe0) has proven to be an effective reactive medium for the remediation of
U(VI) by acting as an efficient electron-donor for the chemical reduction of U(VI) to
U(IV) [4, 5]. The Fe2+/Fe0 redox couple (Eh0’ = -0.44 V) has a lower standard reduction
141
potential than hydrogen gas H+/H2 (Eh0’ = -0.41 V) [6]. Fe0 is thus a thermodynamically
favorable electron donor for the reaction with U(VI) since the standard reduction
potential of the uranyl ion - uraninite redox couple (U(VI) (UO22+/UO2) is much higher
(Eh0’ = +0.41 V) [7]. Furthermore, Fe0 is an insoluble material that can serve as a slow
release electron donor. In this manner, Fe0 can provide a large reservoir of electron
equivalents that can buffer against U(IV) reoxidation in the aquifer upon incidental
exposures to oxidants like dissolved oxygen. These characteristics make Fe0 very
attractive for low-maintenance remediation methods. Fe0 has been tested in laboratory
and field operations for the remediation of uranium [4, 8]. Permeable reactive barriers
(PRB) have become the most common approach for the application of Fe0 for U(VI)
treatment [9, 10].
Past research has suggested that the predominant mechanism for U(VI) removal with
Fe0 is the chemical reductive precipitation forming insoluble U(IV) [4, 11-14]. However,
co-precipitation of U(VI) with oxidized iron corrosion products has also been suggested
to be an important mechanisms in some studies [15, 16]. Nevertheless, these chemical
mechanisms of uranium attenuation may be kinetically limited [12].
Previous works have demonstrated that a large variety of microorganisms can
perform U(VI) reduction with different electron donors [17, 18]. It is also well known
that anoxic corrosion of Fe0 provides cathodic H2 that can be utilized by many
autotrophic microorganisms, including methanogens and sulfate-reducing bacteria [6, 19-
142
21]. However, the role of Fe0 as direct or indirect source of electron equivalents in
biologically catalyzed uranium reduction processes has not yet been explored. Studies
have found significant changes in the microbial population near PRBs with Fe0 treating
groundwater contaminated with uranium, technetium and nitrates [22]. More recent
evidence showed a successful long-term reduction and immobilization of U(VI) in a
sand/ Fe0 packed bed column inoculated with anaerobic granular sludge [23],
demonstrating a significant role of biological activity in stimulating the reduction of
U(VI) by Fe0.
The objective of this study was to demonstrate that the reduction of U(VI) by Fe0 can
be enhanced by the activity of microorganisms in an enrichment culture developed from
the sand/ Fe0 packed bed column. The goal was also to demonstrate that the U(VI)
reduction can be sustained over long time periods
The study also explored the
mechanisms that govern the biological enhancement of U(VI) removal by Fe0 in order to
improve the applicability of Fe0-PRB technology for uranium remediation.
143
5.3. Materials and methods
5.3.1. Basal medium
The mineral medium used in the batch experiments was adapted from that used in
a previous study [23]. The composition of the medium for experiments with sulfate
present (MS) was the following (in mg L-1): NH4HCO3 (5.0), K2HPO4 (2.0), Ca(OH)2
(1.0), yeast extract (1.67) and MgSO4·7H2O (50.0). In experiments without sulfate
(MNS), the MgSO4·7H2O was substituted by MgCl2·7H2O (41.0 mg L-1). In both the MS
and the MNS media, the following final concentration of trace elements was used (in mg
L-1): H3BO3 (0.01), FeSO4·7H2O (0.56), ZnSO4·7H2O (0.02), MnSO4·7H2O (0.08),
(NH4)6Mo7O24·4H2O (0.04), AlK(SO4)2·12H2O (0.04), NiSO4·6H2O (0.02), CoSO4·7H2O
(0.47), Na2SeO3·5H2O (0.02), Na2WO4·H2O (0.10), CuSO4·5H2O (0.03), EDTA (0.20),
and resazurin (0.04). After adjusting the pH to 7.5, the medium was sterilized in an
autoclave (Yamato Scientific America Inc., Santa Clara, CA, USA) at 120°C for 20 min.
After cooling down, the medium was amended with a filter-sterilized NaHCO3 solution to
a final concentration of 1.0 g L-1.
144
5.3.2. Source of inoculum
Enrichment cultures utilized in this study were first inoculated from the effluent
of a continuous column packed with Fe0/sand that reduced U(VI) with elemental iron
(Fe0) as electron donor [23]. These enrichment cultures were developed and maintained
through serial transfer batch experiments as will be explained on the next section, and
served as the source of inoculum for the experiments in this study.
5.3.3. Batch experiments
Batch experiments were performed in 160-mL sterilized serum bottles supplied
with 100 mL of basal medium (either MS or MNS according to the experimental needs).
In the experiments with Fe0, solid Fe0 powder was sterilized for 1 h in an ultraviolet
cross-linker (302 nm, 115V, Cole-Parmer, Vernon Hills, IL, USA) and then added to the
bottles to a final concentration of 56.0 mg L-1. On the other hand, for U(VI) reduction
tests, aliquots were added from a 10 mM U(VI) stock solution to attain final
concentrations of 30, 60 or 100 µM U(VI), depending on the assay. Anaerobic conditions
were achieved by flushing the headspace with a N2/CO2 gas mixture (80/20, v/v) [18].
For biological treatments, aliquots corresponding to 5% v/v (unless otherwise indicated)
of planktonic inoculum from an active enrichment culture were added to each bottle
inside an anaerobic glove box (COY Laboratory Products Inc., Grass Lake, MI, USA)
145
after the flushing described, re-flushed with N2/CO2 gas mixture after inoculation and
before incubation. The pH of the batch experiments with Fe0 was among 6.6 – 6.8. All
assays were conducted in duplicates. All treatments and controls were maintained in the
dark under static incubation at 30°C.
5.3.3.1. Enrichment process of U(VI)-reducing/ZVI-oxidizing culture
The microbial enrichment was carried out in batch experiments (as specified
above) with MS, 56.0 mg Fe0 L-1, and 30 µM U(VI). An aliquot of a previously active
enrichment (9-10% v/v) was used for transfer 0 through 3, and was decreased to 5% v/v
from transfer 4 onwards (see Table 5.1). During each transfer, the enrichment culture
received an initial feeding of 30 µM U(VI) followed by 3 to 4 respikes to restore the
same U(VI) concentration each time soluble U was consumed. The N2/CO2 gas mixture
was replenished at the end of each respike in the transfers to maintain the anaerobic
conditions. Controls for each transfer consisted of treatments that were not inoculated
(Fe0 abiotic control), as well as treatments that were inoculated but did not receive Fe0
(endogenous control). All controls and treatments were monitored for soluble U over
time during the whole experiment. A new transfer to fresh medium was made after the
complete consumption of the five feedings (initial and four respikes).
146
5.3.3.2. Uranium reoxidation
In order to determine if the reductive process was the main mechanism for
uranium removal in a Fe0 treatment inoculated with the enrichment co-culture and respiked three times with U(VI) (30 µM), the reoxidation of uranium was assayed as a
subsequent step by exposition to O2. The reoxidation of uranium was assayed in biotic
incubations (inoculated and with Fe0) after aqueous phase uranium removal had taken
place from all respikes, after approximately 40 days of incubation. The O2 was supplied
to the bottles by flushing the headspace (60 mL) with a gas mixture of He/CO2/O2
(80:20:20 v/v) by 5 min on day 40. The resolubilization of uranium was monitored over
time with ICP-OES as is indicated in analytical techniques section of the paper. To
ensure that the O2 did not become limiting, the bottles were flushed with same gas
mixture after each sampling of uranium analysis.
5.3.3.3. Experiment with different electron donors
A batch experiment was conducted as described above to test three different
electron donors (e-donors) for U(VI) reduction, namely, Fe(II), ethanol, and H2. All edonors were added in a final concentration of 2 meq L-1, equivalent to the amount of
electron donor added in Fe0 experiments. Fe(II) was supplied from a 6 mM FeCl2·4H2O
solution, ethanol from a 0.5 mM solution, and H2 from a H2/CO2 (80:20, v/v) gas
147
mixture. The following set of treatments and controls were prepared: no inoculum, no
electron-donor; endogenous (inoculum only); abiotic with Fe(II); abiotic with H2; abiotic
with ethanol; inoculum with Fe(II); inoculum with H2, and inoculum with ethanol. These
treatments and controls were monitored for soluble U over time.
An additional experiment was conducted to test the biotic and abiotic U(VI)reduction by Fe(II) (1 mM) at different pH values (6.4 – 7.2). The pH range was created
by varying the concentration of CO2 in the headspace (0, 5, 15 and 20% CO2).
5.3.3.4. Preincubation experiments
In order to better understand the enhancement in the rate of microbial U(VI)
reduction by Fe0, a batch experiment was conducted to compare the reduction of U(VI) in
treatments containing Fe0 previously exposed to the enrichment culture and that of
treatments supplied with fresh Fe0 and inoculum. The media tested (both MS and MNS)
was amended with Fe0. The following pre-incubated treatments and controls were set-up:
Fe0 abiotically-preincubated with medium, a control pre-incubated with inoculum-only,
and the full treatment in which Fe0 was preincubated biologically in the medium with the
enrichment culture. The pre-incubations were incubated during 31 days, applying a
concentration of 100 µM U(VI) at the end of the pre-incubation period to start the
experiment. The following non-pre-incubated treatments and controls were set up fresh at
148
the end of the pre-incubation period: abiotic Fe0, endogenous control and inoculum with
Fe0. A concentration of 100 µM U(VI) were also supplied to the fresh treatment and
controls. Soluble Fe concentration (Fe(II) and Fe(III)) were measured during the preincubation stage and during the actual experiment after the pre-incubation. Soluble U
concentration was analyzed over time after its addition to the experiments.
5.3.3.5. Production of H2 by corrosion of Fe0
In order to study the role of enrichment culture on cathodic H2 production from
solid phase Fe0, a batch experiment was set-up with media MS in the absence of U
consisting in Fe0 abiotic controls and the full biological treatment with Fe0 inoculated
with the enrichment culture. H2 gas concentration in the headspace was measured over
time.
5.3.3.6. H2 consumption in presence of ferric hydroxide and magnetite
The biological consumption of H2 was studied in experiments with different types
of solid Fe(III)-containing minerals in MNS medium. The treatments and controls
included: abiotic medium control with H2, Fe(III)-mineral abiotic control with H2, an
inoculum only control, inoculum with Fe(III)-mineral control, and the full biological
149
treatment with Fe(III)-mineral, and H2 inoculated with enrichment culture. After the
anaerobic flushing as described above, the bottles were taken for inoculation and
application of iron compounds to the anaerobic chamber. Magnetite and ferrihydrite were
applied to the corresponding bottles to a final concentration of 2 mM Fe(III). After
replenishing the headspace with N2/CO2, 4.0-mL aliquots of a mixture H2/CO2 (80:20,
v/v) were applied to the headspace of the bottles to arrange a final H2 concentration of
5.3%, which is equivalent to 1.3 mmol H2 Lliq-1.
5.3.3.7. Sulfate reduction by the enrichment culture
In order to evaluate whether sulfate (SO42-) could be a sole electron acceptor for
the enrichment culture, a batch experiment was conducted with H2 serving as the electron
donor and SO42- as electron acceptor (in the absence of iron). The basal medium used
was MNS, and inoculum was added as previously mentioned in manuscript. SO42- was
added as a 1-mL aliquot from a concentrated Na2SO4 stock solution to provide a final
concentration of 0.603 mM SO42- in the treatment bottles. H2 was added from a gas
mixture H2/CO2 (80:20) corresponding a final concentration of 1.207 mmol/Lliq in the
bottles. The following controls and treatments were set up: abiotic with H2, abiotic with
SO42-, abiotic with SO42- and H2, biological with H2 (no SO42-), biological with SO42- (no
H2), and complete biological treatment with SO42- and H2. Anaerobic flushing and
sealing were carried out in the same way as specified in batch experiments.
150
5.3.4. 16S rRNA gene clone libraries
Community genomic DNA was extracted from 5 mL samples taken from the
enrichment at the 11th transfer by a modification of the extraction protocol described by
the manufacturer for Genomic DNA from bacteria (FastDNA Spin Kit for Soil,
Qbiogene, Inc, Carlsbad, CA, USA). Extraction blanks were processed in parallel
throughout the full procedure as negative controls to evaluate potential DNA
contamination from reagents. The 16S rRNA gene was PCR amplified from community
DNA extracts using universal primers 27F and 1492R [24]. To verify the integrity of the
amplification, both positive, negative (no template DNA) and original inoculum reactions
were included. The PCR products were purified using PureLinkTM Quick Plasmid
Miniprep Kit (Invitrogen, Carlsbad, CA, USA) according to the protocol described by the
manufacturer. The purified PCR products were checked on a 1.5% agarose gel. The
identities of amplicons were also confirmed by verifying the molecular size of the
amplicons on the gel. The purified PCR products were cloned into plasmid vector
pCR®2.1-TOPO® using the TOPO TA cloning system (Invitrogen) according to the
protocol described by the manufacturer. The procedure of 16S rRNA gene based clone
library analysis protocol and sequences used in this work are described in details by Sun
[25]. Selected clones representing each phylotype obtained in the enrichment have been
deposited in the GenBank database with accession numbers:
Stenotrophomonas sp.
enrichment culture clone U-co-culture-1, JF729003 and Dechloromonas sp. enrichment
culture clone U-co-culture-2, JF729004. Sequence data were aligned with ClustalX
151
including 16S rRNA gene sequences from reference bacterial strains (GenBank) and
unique phylotypes recovered from the enrichment, and a tree was constructed using
PAUP* version 4.0b10 employed with the neighbor-joining algorithm [26]. The
GenBank accession numbers for the sequences used to prepare Figure 5.3 are as follows:
Stenotrophomonas sp. TS23, EU073089; uncultured bacterium clone R2-B12, FJ971741;
uncultured bacterium clone ctg1_TOPO1-17, EU708503; Dechloromonas sp. JM,
AF323489; Dechloromonas agitatus strain CKB, AF047462; perchlorate-reducing
bacterium EAB2, AY265879; uncultured beta proteobacterium A2-4c17, EU236249;
uncultured beta proteobacterium Glu+NO3-c24, EU236237; Planctomyces sp.; Schlesner
130; X81952.
5.3.5. Magnetite synthesis procedure
The method utilized to synthesize magnetite was adapted from Jolivet and Tronc
[27] and Missana et al. [28]. MilliQ water for the synthesis was previously boileddegassed in order to attain anaerobic conditions and allowed to cool down to room
temperature. A 100-mL solution containing 39.762 g FeCl2·H2O and 16.7 mL HCl in
degassed MilliQ water, and a 400-mL solution containing 64.884 g FeCl3 in degassed
MilliQ water were prepared in a glove box (COY Laboratory Products Inc., Grass Lake,
MI, USA). The solutions were mixed and stored in a glass bottle. Inside the anaerobic
glove box, an aliquot of 60 mL of this solution was taken to a 100-mL serum bottle, and
152
subsequently sealed with butyl rubber septa and aluminum crimps. A 400-mL solution
was made containing 70.87 mL of NH3 reagent, and added to a ball flask, closing its ports
with rubber septa and leaving it under continuous flushing with N2 gas. Then, 50 mL of
the anaerobic Fe-solution were slowly added by injection to the NH3 solution under
continuous N2 bubbling and vigorously stirring. After the addition and observation of
black precipitates, flushing was stopped and the system was left sitting for a moment.
The closed system containing the synthesized material was taken again to the anaerobic
glove box, in which the product was separated by magnetic settling and decanting.
Spectra/Por® Float-A-Lyzer® G2 cellulose dialysis tubing (Spectrum Labs,
Rancho Dominquez, CA, USA), with an approximate molecular weight cut off (300,000
Daltons) and a nominal volume of 10 mL was used to remove all the salts from the solid
product. In the anaerobic chamber, the dialysis unit was soaked in degassed MilliQ water
for 30 min. Next, the dialysis unit was transferred to a dialysate buffer, which was
prepared by diluting 10 times the anoxic Fe-solution in degassed MilliQ water (the buffer
was used instead of pure MilliQ water to avoid breakage of the membrane by osmotic
pressure). A volume of 500-mL of the same dialysate buffer was added to a glass flask.
Then, 5 mL of the remaining wet solids were carefully loaded inside the dialysis bag; and
the bag was sealed and placed in the dialysate buffer with help of a flotation ring.
Stirring was initiated in order to attain a moderate spinning, taking appropriate care not to
create a strong current. Fresh buffer was replaced after 3, 8 and 14 h, leaving this last
one overnight. After dialysis, the unit was opened and the sample was aspirated slowly
153
and dispensed in a new container, in which it was left to settle and excess water was
decanted under anaerobic conditions. The samples from this separation were stored in
sealed containers in the anaerobic glove box. Solids for characterization by X-ray
X
diffraction (XRD) were appropriately dried under a N2 atmosphere.
Figure 5.1. Results from X
X-ray
ray diffraction (XRD) of synthetic magnetite, with labels
corresponding to the Fe3O4 pattern (JCPDS-ICDD Card #75-16
1609).
154
5.3.6. Analytical methods
5.3.6.1. Soluble U
Liquid samples for uranium analysis were centrifuged in EppendorfTM tubes (1.5mL) at a relative centrifugal force (RCF) of 10,621 x g for 10 min. A volume of 0.5 mL
of the supernatant was taken and preserved into 3.3 mL of a 3% HNO3 solution and 1.2
mL of MilliQ water. Soluble U was measured by using an Inductively Coupled Plasma –
Optical Emission Spectrometry (ICP-OES) system (Optima 2100 DV, Perkin-Elmer,
Shelton, CT, USA) at a wavelength of 385.958 nm. The detection limit for U was 10 µg
L-1.
5.3.6.2. Soluble iron
The concentration of Fe(II) and Fe(III) in filtered samples (0.45 µm) was
analyzed by the colorimetric phenanthroline method [29]. In this method, Fe(III) is
reduced to Fe(II) with 0.1 M hydroxylamine hydrochloride (NH2OH·HCl) solution, and
1,10-phenanthroline is applied as an indicator. The final Fe(II) concentration was
measured at 560 nm using an Agilent 8453 UV-spectrophotometer (Agilent, Palo Alto,
CA, USA). The calculation of Fe(III) present in the sample was made by subtracting the
155
Fe(II) of the non-NH2OH·HCl reduced sample from the total Fe(II) of the NH2OH·HCl
reduced sample.
5.3.6.3. Hydrogen
A gas chromatography system (7890A model, Agilent Technologies, Palo Alto,
CA, USA) was used to measure H2 gas. The system was equipped with a thermal
conductivity detector (TCD), and the column used was a Supelco CarboxenTM 1010 Plot
(fused silica capillary column, 30 m x 0.53 mm, average thickness 30 µm, Supelco,
Bellefonte, PA, USA). The carrier gas was nitrogen (N2) to avoid any interference by
similarity in thermal conductivity of helium. The injection volume was 100 µL. The
temperature of the injection port, column and detector were 220, 100 and 230°C,
respectively.
5.3.6.4. Sequential Extraction
The sequential extraction procedure of U in the solid phase was conducted as
previously reported [18].
156
5.3.6.5. X-ray photoelectron spectroscopy
Preparation of the solid phase from a five-times respiked treatment with 30 µM
U(VI) by anaerobic drying was performed as previously reported by Tapia and coworkers
[18].
All the samples were introduced in the analytical system immediately after
preparation to limit contamination and aging effects, and were compressed against an
indium foil. XPS analyses and high-resolution spectra were performed using an imaging
Kratos Axis Ultra (U.K.) X-ray photoelectron spectrometer equipped a conventional
hemispherical analyzer in spectrum mode. The X-ray source was monochromatized Al
KR operating at 150 W (15 keV/10 mA). A wide compositional survey scan was acquired
using pass energy of 160 eV, followed by high-resolution elemental spectral regions
using pass energy of 20 eV. A charge neutralizer was used at all times. Temperature was
controlled below 0ºC and a low-energy electron flood gun was used to minimize surface
change of the samples due to X-ray beam damage. Data analysis was performed with
Vision Processing data reduction software (Kratos Analytical Ltd.). Peak binding
energies are referred to the carbon C 1s peak at 284.6 (aromatic carbon), which is
considered as a calibrant to adjust the binding energy scale for all reported binding
energies. Standards of UO3, UO2Cl2·H2O and UO2 were analyzed as a reference of the
oxidation states of uranium in the sample.
157
5.3.6.6. Other determinations
Measurements of pH were performed in a VWR SympHony SB20 electrode
according to Standard Methods [29].
5.3.7. Chemicals
Uranium (VI) in the forms of uranyl chloride trihydrate (UO2Cl2·3H2O) and
uranium trioxide (UO3), as well as uranium (IV) in the form of uranium dioxide (UO2)
were all purchased from International Bio-Analytical Industries Inc. (Boca Raton, FL,
USA). Powder Fe0 (< 10 µm, >99.9% purity), ethanol (99.5%), potassium permanganate
(KMnO4, 99.0+%) and hydroxylamine hydrochloride (NH2OH·HCl, 99%) were
purchased from Sigma Aldrich Co. (St. Louis, MO, USA). Magnetite (Fe3O4) was
synthesized by the method described previously. Iron (III) hydroxide (Fe(OH)3) slurry
(13%) was purchased from NOAH Technologies Corporation (San Antonio, TX); the
material was 99.5% pure. Ammonium bicarbonate (NH4HCO3, 21.3-21.7% as NH4+),
ferrous chloride tetrahydrate (FeCl2·4H2O, 99.0+%), sodium hydroxide (NaOH, 97.0+%),
nitric acid (HNO3, 70.0%), ammonium acetate (NH4C2H3O2, 97.0+%), 1,10phenanthroline monohydrate (C12H8N2·H2O) and hydrochloric acid (HCl, 37.0%), were
all supplied by Fisher Chemical (Fair Lawn, NJ, USA). Magnesium sulfate
(MgSO4·7H2O), calcium hydroxide (Ca(OH)2) and potassium phosphate dibasic
158
(K2HPO4, 99.0+%) were purchased from J.T. Baker (Phillipsburg, NJ, USA). Yeast
extract was supplied from BD (Sparks, MD, USA). Sodium bicarbonate was obtained
from Pfaltz & Bauer (Waterbury, CT, USA).
5.4. Results
5.4.1. Reduction of U(VI) with enrichment culture
An enrichment culture (EC), originating from a U(VI) reducing biological column
with Fe0, displayed activity in stimulating U(VI) reduction with Fe0 beyond abiotic rates
for 20 transfers over the course of 28 months. Each transfer of several 30 µM spikes
lasted approximately 40 days. A summary of the biological and abiotic rates of U
removal during 20 transfers is shown in Table 5.1. Throughout the maintenance of the
enrichment culture, the initial rate of U removal in biologically active Fe0 containing
cultures was 3.0- to 27.4-fold faster than in the parallel run abiotic controls. In order to
gain knowledge on the identity of the microbial community, the enrichment culture was
characterized with a clone library.
159
Table 5.1. Rates of uranium reduction over 20 transfers of the enrichment culture at 30
µM U(VI).
Zero order rate (µ
µM U(VI) d-1)
Transfer
% Inocula* Initial
Spike 1
Spike 2 Spike 3 Spike 4 Abiotic
0
10
1.25
2.57
2.28
2.83
3.09
0.10
1
9
3.06
4.78
9.95
8.21
N/A
0.70
2
10
2.43
8.29
8.90
8.90
N/A
0.70
3
10
3.76
8.19
5.86
6.04
12.66
0.30
4
5
2.77
8.21
9.11
8.40
8.15
0.70
5
5
2.57
7.65
10.65
8.40
13.37
0.70
6
5
2.67
9.78
9.80
5.46
8.41
0.30
7
5
2.74
5.54
6.52
5.27
1.69
0.10
8
5
2.72
4.80
5.77
3.88
4.10
0.75
9
5
2.78
4.93
5.88
4.59
4.16
0.29
10
5
3.00
5.59
3.67
3.93
4.24
0.37
11
5
2.39
4.59
9.15
7.58
7.90
0.22
12
5
2.81
3.66
3.16
3.10
3.03
0.03
13
5
2.14
7.70
2.89
8.30
7.28
0.50
14
5
2.08
2.74
3.10
3.80
3.55
0.66
15
5
5.28
1.84
3.59
6.26
3.72
0.47
16
5
2.07
2.96
2.78
3.76
N/A
0.56
17
5
2.29
4.15
4.99
5.58
3.95
0.75
18
5
4.77
5.37
9.95
5.19
7.15
0.48
19
5
2.57
4.21
4.85
3.56
3.00
0.57
20
5
2.62
3.95
3.11
2.24
6.61
0.45
* Supernatant from the full treatment of previous transfer.
** From transfer 5 onwards, media was supplemented with 5% YE.
160
5.4.2. Microbial community composition of the enrichment culture
The microbial community composition in the established enrichment was
analyzed by preparing a 16S rRNA gene clone library with primers targeting bacteria.
Rarefaction analysis for the clone library (Figure 5.2) suggested that 12 clones
predominated in the community composition of the clone library. The 12 clones analyzed
in this study fell into two phylogenetic divisions, β-Proteobacteria and γ-Proteobacteria,
accounting for 33.4%, and 66.6% of all the clones as shown in Figure 5.3. The
predominant PCR-amplified clones of the enrichment culture contained only 2 bacterial
ribotypes, one within the genus Dechloromonas and the other within the genus
Stenotrophomonas.
Number of unique clones
detected
2.5
2
1.5
1
0.5
0
0
5
10
Number of clones tested
15
Figure 5.2. Evaluation of representative clones obtained from the enrichment culture by
rarefaction analysis.
161
Figure 5.3. Phylogenetic tree for the bacteria identified in the co-culture enrichment
with the universal bacteria PCR primer set 27F and 1492R.
5.4.3. U(VI) transformation
A sequential extraction protocol was utilized to characterize the mechanism of the
U removal in the EC. The sequential extraction was performed on the solids deposited in
the enrichment culture after five consecutive spikes with 30 µM U(VI). The sequential
extractions consisting of MilliQ water, 1 M NaHCO3 and 1 M HNO3 are intended to
distinguish between water soluble U, adsorbed U(VI) and reductively precipitated U(IV),
respectively. The total U extracted corresponded to 14.7 µmol, which accounted for 98%
of the aqueous U(VI) fed and removed by the culture. Figure 5.4 shows the different
fractions of U recovered. The water-soluble and adsorbed U(VI) were very low. The
162
overwhelming majority of the extracted U (94.7%) corresponded to U(IV) suggesting that
reductive precipitation was the main mechanism of U removal.
Figure 5.4. Distribution of recovered mass of U in a treatment with 5 feedings of 30 µM
U(VI) through sequential extraction with H2O (soluble), HCO3 (adsorbed) and HNO3
(reduced).
To confirm these results, material deposited in a similar EC was exposed to O2 at
the end of an experiment. Results for this experiment are sh
shown
own in Figure 5.5. The O2
oxidized the deposited U(IV) causing most of the biogenically immobilized U to become
solubilized. Complete reoxidation was achieved after 8 days of O2 exposition. The results
on chemical reoxidation provide additional evidence th
that
at the biological removal of
163
uranium was mediated by the reduction of the soluble hexavalent uranium (U(VI)) to its
insoluble tetravalent form (U(IV)).
Figure 5.5. Time-course
course of uranium reduction by U
U-enrichment
enrichment culture added with Fe0 as
electron donor and subsequent uranium reoxidation using O2 as oxidant. The dotted
horizontal line indicate the total amount of U(VI) added.
ray diffraction (XRD) and X
X-ray
ray photoelectron spectroscopy (XPS) were used
X-ray
to obtain more evidence of the U speciation after incubation with the enrichment culture.
XRD only provided a weak signal for uraninite in the samples (Figure 5.6).
.6).
164
Figure 5.6. Results from X
X-ray
ray diffraction (XRD) of U(IV). (A) XRD of uraninite
UO2(s) standard. (B) XRD pattern of a solid sample from a treatment with Fe0 and
enrichment culture after the complete consumption of eight consecutive feedings of 60
µM of U(VI). Labels in both panels correspond to the UO2 pattern (JCPDS-ICDD
(JCPDS
Card
#73-1715).
165
However, further evidence obtained by XPS (Figure 5.7A) revealed that one of
the runs for the sample had U4f 7/2 and U4f 5/2 peaks with binding energies of 379.2 and
390.1, respectively. A second run of the sample (Figure 5.7B) presented values of 379.8
and 390.7 for these peaks. The spectra of U(VI) standards showed U4f 7/2 and U4f 5/2 peak
binding energies at 376.5 and 387.5 for UO2Cl2 (used in the batch experiments) and
377.0 and 388.0 for UO3.
The standard of U(IV) tested was in the form of UO2
(uraninite), with higher binding energies located at 380.5 and 391.5, respectively.
Results indicate that the oxidation state of the sample was predominantly U(IV). A small
amount of U(VI) in the first run of the sample may account for the slightly lower values
of the U4f 7/2 and U4f 5/2 peak binding energies. This could explain the very amorphous
phase found in XRD, indicating that U(VI) might be in other forms.
166
Figure 5.7. X-ray
ray photoelectron spectroscopy (XPS) U4f7/2 and U4f5/2 binding energy
spectra for two replicates of the solids from the enrichment culture with Fe0 respiked 5
times with 30 µM of U(VI) (pink line in panels A and B), as well as U(IV) in the form of
UO2 (blue line), U(VI) in the forms of UO2Cl2 (red line) and UO3 (black line).
167
The EC were tested to study the impact of U(VI) concentration on the U removal
pattern. Figure 5.8 shows the time course of soluble U(VI) removal in the presence of the
enrichment-culture and Fe0 at two uranium concentrations. At an initial U(VI)
concentration of 30 µM (Figure 5.8A), soluble U(VI) was removed 4.0-fold faster in the
presence of the enrichment culture compared to the abiotic control with Fe0 but lacking
inoculum. The U removal started immediately without any lag-phase. Neither the
biological controls (endogenous control) nor abiotic controls without Fe0 demonstrated
any noteworthy U(VI) reduction during the whole period of the experiment (40 days). At
an initial U(VI) concentration of 60 µM (Figure 5.8B), the U removal in the inoculated
Fe0 treatment started after a 10 d lag phase. Once the U removal started it occurred at a
rate approximately 4.8-fold faster than the abiotic Fe0 control. The abiotic rates of U(VI)
reduction also increased with increasing U(VI) concentrations with rates of 0.81, 1.43
and 4.67 µM d-1 at an initial U concentration of 30, 60 and 100 µM respectively.
168
Figure 5.8. Experiment with two concentrations of uranium. Panel A) 30 µM and B) 60
µM U(VI). Legends: -----◊---, Abiotic, no Fe0; ---□---, Biological, no Fe0; —○—, Abiotic
with Fe0; —●—, Biological with Fe0.
169
The increase in concentration from 30 to 60 µM had the effect of creating a lag
phase. The biotic rate after the lag phase increased approximately two-fold with the
higher U(VI) concentration. The experiments of Figure 5.8 also clearly demonstrate that
the Fe0 enabled biologically stimulated U(VI)-reduction compared to the lack of any
significant reduction in endogenous controls, suggesting a potential role of Fe0 as an
electron donor to the EC.
5.4.4. Alternative electron donors for U(VI) reduction
The possibility that other reduced compounds aside from Fe0 could serve as
electron donors to the EC was considered. An experiment was conducted with Fe2+,
ethanol, and H2 as electron donors, with results shown in Figure 5.9. These results
demonstrated that they do not have any effect on U(VI) reduction. These alternative
compounds were not electron donors supporting U(VI)-reduction by the enrichment
culture, even though two of the compounds tested (Fe2+ and H2) are major anoxic
corrosion products of Fe0.
170
Figure 5.9. Time course of uranium with different electron donors in the presence of
microbial co-culture
culture for 30 days of incubation. Legends: ------,, No inocula, no electron
donor; — —, Endogenous; ---▲---, Abiotic + Fe(II) (pH 6.48); ---■---,, Abiotic + H2; --●---,, Abiotic + Ethanol; —▲—, Biological + Fe(II) (pH 6.48); —■—,, Biological + H2;
—●—, Biological + Ethanol.
Fe2+ was also tested for the biotic and abiotic reduction of U(VI) at different pH
values, and results demonstrated that at the pH of the EC no abiotic nor biotic removal of
aqueous U(VI) occurs. The findings merit consideration of alternative hypotheses
hypothese for the
role of Fe0.
171
5.4.5. U(VI) reduction by Fe0 preincubated with the enrichment culture
To gain insight on alternative mechanisms, an experiment was conducted with a
31-day pre-incubation of Fe0 with the EC prior to the addition of 100 µM of U(VI). The
results are shown in Figures 5.10 and 5.11 corresponding to the use of culture medium
without sulfate (MNS) and with sulfate (MS), respectively. The U time courses in both
experiments (Figures 5.10A and 5.11A) reveal that there are large differences in behavior
depending in whether the Fe0 was preincubated abiotically or biologically with the
enrichment culture. The Fe0 preincubated abiotically with the media became non-reactive
towards U as evidenced by the lack of any U removal, regardless of whether the
treatments were inoculated afterwards or not with the EC. This observation was
particularly interesting since fresh (non-preincubated) Fe0 was effective in removing U
both with and without inoculation. These findings suggest that passivation of Fe0
occurred during the abiotic preincubation with the media. On the other hand, biological
preincubation of the Fe0 with the EC increased the initial reactivity of the Fe0 towards U,
removing soluble U even faster than that observed with fresh Fe0. The effect was
observed in both types of media; however, the impact was more pronounced in MS
(Figure 5.11A). During the first 4 days, the biologically pre-corroded Fe0 removed U at a
1.8- and 2.4-fold faster rate than with fresh Fe0 either incubated with the EC or
abiotically, respectively. Thus biological preincubation of Fe0 increased the reactivity of
the Fe0 towards U removal. The results also show that the U removal trends observed
were not impacted by the presence or absence of sulfate in the medium.
172
Figure 5.10. Time course for uranium ((panel A) and Fe2+ (panel B)) in treatments
preincubated with MNS media. Legends: —◊—, Abiotic, no Fe0; —▲—
— Preincubated
inocula; ---▲---, Non-preincubated
preincubated inocula; —□—, Abiotically-preincubated
preincubated Fe0; ---□---,
Non-preincubated Fe0; —
—■—, Abiotically-preincubated Fe0 + non-preincubated
preincubated inocula;
—●—,, Preincubated Fe0 / inocula; ---●---, Non-preincubated Fe0 + non-preincubated
non
inocula.
173
Figure 5.11. Time course for uranium ((panel A) and Fe2+ (panel B)) in treatments
treatme
preincubated with MS media. Legends: —◊—, Abiotic, no Fe0; —▲—
— Preincubated
inocula; ---▲---, Non-preincubated
preincubated inocula; —□—, Abiotically-preincubated
preincubated Fe0; ---□---,
Non-preincubated Fe0; —
—■—, Abiotically-preincubated Fe0 + non-preincubated
preincubated inocula;
—●—,, Preincubated Fe0 / inocula; ---●---, Non-preincubated Fe0 + non-preincubated
non
inocula.
174
Treatments were also conducted by incubating 100 µM U(VI) with fresh Fe0.
Initially there is little difference in the U removal rate between the inoculated and abiotic
treatments. However after a lag phase of approximately 7 days, the U removal in the
inoculated treatment increased significantly up to approximately 2-fold more than the
abiotic rate (Figures 5.10A and 5.11A). This behavior was also seen with the 60 µM U
experiment (Figure 5.8B).
During the 31-day pre-incubation of the Fe0 in the absence of U, approximately
22 to 28 mg L-1 of Fe2+ was released regardless of whether the treatments were inoculated
or not (data not shown). The fate of Fe0 during the U exposition period is shown in Figure
5.10B in the sulfate free medium (MNS). The abiotic corrosion of fresh Fe0 to Fe2+
occurred rapidly over the first 15 days; thereafter it slowed down and subsequently
ceased completely by day 25.
About half of the Fe0 was recovered as Fe2+. The
biocorrosion of fresh Fe0 started similarly as the abiotic corrosion, but the rapid release
Fe2+ continued beyond day 15. After 35 days, 80% of the Fe0 was converted to Fe2+. By
comparing the two Fe2+ time courses, the fresh Fe0 in the abiotic experiment started to
become passivated around day 15. Measurements of soluble iron detected very low
quantities of Fe3+.
This passivation effect is also evident from the lack of any additional Fe2+ release
in Fe0 samples that were pre-incubated abiotically regardless if the samples were
incubated biotically or abiotically in the subsequent U exposure period. In stark contrast,
175
samples pre-incubated biotically, continued to release Fe2+ during the subsequent U
exposure period. The data suggest the biocorrosion of Fe2+ continued in the biotically
pre-incubated samples.
The iron corrosion patterns were similar with the sulfate-containing medium
(Figure 5.11B) except that the maximum release of Fe2+ in biotically incubated
treatments was distinctly less after day 15. The lowered release of Fe2+ is most likely due
to the onset of sulfate reduction and precipitation of iron sulfides as a result.
The EC was tested for its ability to reduce SO42- to H2S with H2. Figures 5.12,
5.13 and 5.14 show the time course of H2, SO42- and [H2S]liq in this experiment,
respectively. Only after a lag phase of approximately 16 days, the full treatment with
SO42- and H2 in the presence of the enrichment culture was able to totally consume the
amended H2 in contrast to the controls where the H2 remained almost constant during the
whole experiment (Figure 5.12). Results demonstrated that 4.81% of H2 in the headspace
gas were consumed in the experiment, corresponding to 1.16 mmol Lliq-1 (2.32 meq e- L1
). Consumption of SO42- was only detected at significant levels in the full biological
treatment (Figure 5.13). This consumption started to occur at approximately day 16. A
total of 20.4 mg L-1 of SO42- (0.21 mM SO42-) was consumed, corresponding to 73.3% of
the theoretical amount of SO42- from the observed H2 consumption. A steep formation of
H2S after day 16 was only observed in the full biological treatment (Figure 5.14),
concurrent with the start of H2 consumption. This production reached a maximum at day
176
33. At that time, 5.9 mg L-1 of [H2S]liq had accumulated, representing a 91.1% of the
expected [H2S]liq from the amount of SO42- formed. The experiment demonstrated that
was enrichment of bacterial species capable of sulfate reduction. This observation
observati led to
the choice of a non-sulfate
sulfate containing medium for the experiments of Fe(III) reduction by
H2.
Figure 5.12. Time course of H2 consumption by the enrichment co-culture
culture with SO42- as
sole electron acceptor. Legends: ------, Abiotic + H2 (no SO42-); ---●
●---, Abiotic +
SO42- + H2; —○—, Biological + H2 (no SO42-); —●—,, Biological + SO42- + H2; ---■---,
Abiotic + SO42- (no H2); —■—, Biological + SO42- (no H2).
177
Figure 5.13. Time course of SO42- consumption by the enrichment co-culture
culture with H2 as
electron donor. Legends: ------, Abiotic + H2 (no SO42-); ---●---,, Abiotic + SO42- + H2;
—○—, Biological + H2 (no SO42-); —●—, Biological + SO42- + H2; ---■---,
--Abiotic +
SO42-- (no H2); —■—, Biological + SO42- (no H2).
178
Figure 5.14. Time course of H2S production by the enrichment co-culture
culture with SO42- as
sole electron acceptor and H2 as electron donor. Legends: ------,, Abiotic + H2 (no
SO42-); ---●---,, Abiotic + SO42- + H2; —○—, Biological + H2 (no SO42-); —●—,
Biological + SO42- + H2; ---■---, Abiotic + SO42- (no H2); —■—,, Biological + SO42- (no
H2).
179
5.4.6. H2 production during anoxic corrosion of Fe0
An experiment was carried out to monitor the formation of cathodic H2 from the
corrosion of Fe0 in order to better understand the mechanism implicated in the
transformation of Fe0 by the EC. Figure 5.15 displays the evolution of H2 during the
abiotic and biotic corrosion of Fe0 under anoxic conditions in the absence of added
U(VI).
The theoretical maximum levels of H2 are represented by horizontal lines
assuming cases in which all the added Fe0 would be converted to soluble Fe2+ or
magnetite (Fe3O4). The H2 evolution during the abiotic corrosion of 1 mM Fe0 reached
72% of theoretical maximum extent (for corrosion to Fe2+) in a matter of 6 days. On the
other hand, although the biotic corrosion of Fe0 had a similar initial rate of H2 evolution,
the evolution peaked on day 4 and the peak H2 was 6-fold less than that of the abiotic
incubation. After day 4, the H2 was completely consumed to negligible levels by day 6.
This observation suggests the enrichment culture was responsible for consuming cathodic
H2 released by corrosion of Fe0.
180
Figure 5.15. Production of H2 during the abiotic (chemical) and biotic anoxic corrosion
of ZVI (100 µM)
M) in the absence of U(VI). Legend: —■—, Fe0 + inocula; —●—, Fe0
without inocula.
5.4.7. Use of H2 as an electron donor to reduce Fe(III) by the enrichment culture
To further identify
entify the role of the EC on the corrosion of Fe0, an experiment was
carried out to demonstrate that Fe(III)
Fe(III)-containing
containing minerals can serve as an electron
acceptor for H2-consumption.
consumption.
Different types of Fe(III)
Fe(III)-containing
containing minerals are
suspected of being corrosion
rrosion products of Fe0. Figure 5.16
.16 shows the consumption of
added H2 (panel A) and production of Fe2+ (panel B) by the enrichment culture in the
181
presence of magnetite (Fe3O4).
Figure 5.17 corresponds to a similar experiment;
however, instead of magnetite, ferrihydrite (Fe(OH)3) is used as the electron acceptor. In
both figures, effective consumption of H2 and corresponding production of Fe2+ is only
observed in the biologically active treatments with either magnetite or ferrihydrite. In
contrast there is no activity in the abiotic controls with either magnetite or ferrihydrite
present nor in the biological control without the Fe(III)-containing minerals or without
added H2. A lag phase of approximately 13 days was observed prior to rapid H2
consumption with magnetite as the electron acceptor. Likewise a similar rapid increase in
the Fe2+ concentration also was observed to commence at day 13 (Figure 5.16B). No lag
phase occurred for the consumption of H2 when ferrihydrite was used as the electron
acceptor. However, with ferrihydrite, there was an 8 d delay before there was a reliable
measurement of the increase in the Fe2+ concentration.
182
Figure 5.16. (A) Use of H2 by the co-culture
culture in the presence of magnetite (Fe3O4). (B)
Time course of soluble Fe2+ concentration. Legends: ---♦---,, Abiotic + H2; ---▲---,
Abiotic + Fe3O4+ H2; —■
■—, Biological + H2 (no iron); —●—,, Biological + Fe3O4 + H2;
---○---,, Biological (no iron, no H2); —○—, Biological + Fe3O4 (no H2); --- ---, Abiotic +
Fe3O4 (no H2).
183
culture in the presence of ferric hydroxide
Figure 5.17. (A) Use of H2 by the co-culture
(Fe(OH)3). (B) Time course of soluble Fe2+ concentration. Legends: ---♦---,
--Abiotic +
H2; ---▲---,, Abiotic + Fe(OH)3 + H2; —■—, Biological + H2 (no iron); —●—,
Biological + Fe(OH)3 + H2; ---○---, Biological (no iron, no H2); —○—
—, Biological +
Fe(OH)3 (no H2); --- ---, Abiotic + Fe(OH)3 (no H2).
184
5.5. Discussion
5.5.1. Microbial enhancement of uranium removal by Fe0
Based on the evidence presented, the enrichment culture was shown to
significantly enhance the rate of U removal by Fe0 compared to abiotic incubations. The
removal of U by the enrichment culture was also shown to be completely dependent on
the presence of Fe0. Different hypotheses of microbial enhancement were considered.
These include direct microbial reduction of U(VI) to insoluble U(IV) by microorganisms
using either Fe0 or corrosion products as electron-donor and U(VI) as the electronacceptor. Another potential hypothesis is that the microbial culture accelerates the
corrosion of Fe0. This could result in the formation biogenic secondary minerals that are
more reactive in chemically reducing U(VI) compared to fresh Fe0 or minerals formed
during abiotic anoxic incubations. A third hypothesis is the enrichment culture prevents
or decreases passivation of the Fe0 surfaces during anoxic incubation. And lastly a final
hypothesis is that the combined formation of Fe2+ and alkalinity during the biocorrosion
of Fe0 creates thermodynamically favorable conditions for Fe2+ to directly reduce U(VI).
All of the hypotheses attempt to explain the accelerated reduction of soluble U(VI) to
solid phase species containing U(IV) by Fe0 in presence of the EC.
185
5.5.2. Reduction of U(VI) as main mechanism of removal
The U removed from solution by the EC with Fe0 was completely recovered by
sequential extraction of the solid residue. Only the oxidative step (with nitric acid) of the
sequential extraction procedure was responsible for most of the U extraction. This
observation is an indication that the U in the solid residue was composed of U(IV) [30,
31]. Chemical oxidation of the culture bottles with O2 completely solubilized the
immobilized U providing an alternative indication of the occurrence of U(IV). in the solid
phase. O2 is well known for its ability to oxidize biogenic U(IV) [32-34]. Lastly, XPS
spectra of the solid residue demonstrated of U(IV) was the main species of U with only a
minor amount of U(VI).
The reduction of U(VI) to U(IV) is an important mechanism during the abiotic
reaction of U(VI) with Fe0 [11-14]. Based on sequential extraction results, the
observations reported from the previous biologically active sand-Fe0 packed bed column
experiments [23] also showed reduction of U(VI) to U(IV) as the predominant
mechanism. Alternative uranium removal mechanisms proposed during reaction of Fe0
with U(VI) involve the adsorption of U(VI) by iron oxides [12, 16]. However, these
alternative mechanisms are not consistent with the release of soluble uranium during
oxidation of the solid phase. Nor are they consistent with a predominance of U(IV)
observed with XPS. Therefore reductive precipitation forming a predominantly U(IV)
containing solid phase is the mechanism most consistent with the data. Secondary Fe(II)
186
minerals formed from Fe0 corrosion have also shown to reduce U(VI), including
magnetite, carbonate green rusts and siderite [15, 35-40]. Formation of mixed valent solid
phases of U(VI)/U(IV) has been observed during the reaction of U(VI) with sulfurminerals, such as pyrite (iron disulfide) [41-43], mackinawite (iron sulfide) [44] and
amorphous ferrous sulfides [45]. This latter evidence could be a reason of the minor
amount of U(VI) found in this study in solid samples from experiments where MS
medium was used.
5.5.3. No evidence for the direct reduction of U(VI) by enrichment culture
H2 and Fe2+ are important anoxic corrosion products of Fe0 [46]. There is ample
evidence that bacteria can utilize H2 to support U(VI) reduction [18, 47-49]. Results from
the test with these potential electron donors indicate that the microorganisms in the
enrichment culture were unable to use anodic Fe(II) or cathodic H2 emanating from Fe0
corrosion to support U(VI) reduction. Therefore there was no evidence that bulk electron
donors originating from Fe0 corrosion served as electron donors for the microbially
catalyzed reduction of U(VI).
The chemical reduction of U(VI) with Fe2+ has been proven possible at mildly
alkaline pH conditions [50]. However, the pH ranges in this study with 20% CO2 and 1 g
L-1 NaHCO3 corresponded to 6.6 – 6.8 and 6.4 – 6.5 for the experiments with Fe0 and
187
Fe2+, respectively, in which the thermodynamic conditions for chemical U(VI) reduction
by Fe2+ are not favorable [50, 51]. This was confirmed by the abiotic incubation of U(VI)
with Fe2+, which resulted in no reaction between U(VI) and Fe2+. This suggests that Fe2+
formed from Fe0 corrosion was not implicated in abiotic U(VI) reduction. Alternative
hypotheses are necessary to explain the enhancement of the U(VI) reduction.
5.5.4. Impact of enrichment culture on Fe0
The role of the EC may be to impact the biocorrosion of Fe0. An alteration of
mineral products formed from the corrosion in turn may have been responsible for the
enhanced kinetics of U reduction. The fact that pre-incubating the Fe0 with the
enrichment culture increased the rate of uranium removal compared to fresh Fe0, clearly
suggests a change in reactivity of iron during the pre-incubation period. Secondary
minerals or soluble corrosion products formed during biocorrosion may be more reactive
with U(VI) compared to fresh Fe0 or passivated iron from abiotic corrosion.
Several types of secondary minerals are known to form during the anoxic
corrosion of Fe0. Examples of secondary minerals formed by abiotic corrosion include
magnetite [15] and carbonate green rusts [15, 52]. In abiotic systems with sulfate, also
mackinawite (ferrous sulfide) [53] has been reported. Some minerals have also been
observed to form either in Fe0-PRBs or in microbial systems with Fe0, including
188
magnetite [54, 55], carbonate green rusts [55], chukanovite (ferrous hydroxy carbonate)
[56], siderite (ferrous carbonate) [54, 55], amorphous ferrous sulfide [54], mackinawite
[54] and greigite (Fe3S4) [23]. Likewise, siderite, chukanovite and vivianite (ferrous
phosphate) have also been observed as products from the biological reduction of
magnetite [57, 58].
5.5.5. Reactive secondary minerals as reductants of U(VI)
Past evidence has shown that Fe(II) or mixed valent Fe(II)/Fe(III) minerals
formed during Fe0 corrosion are responsible for the abiotic reduction of U(VI). For
example, previous studies have shown that magnetite [15, 35, 36, 59] and carbonate
green rusts [15, 37-39] can adsorb and reduce U(VI) to U(IV). However, El Aamrani et
al. [59] have also reported the formation of mixed phases of U(VI)/U(IV) in the presence
of magnetite.
Through similar mechanisms, the sulfur-minerals pyrite [41-43],
mackinawite [44] and amorphous FeS [45] can lead to mixed valent U solid phases
(U(VI)/U(IV)). Siderite was also shown to adsorb and reduce U(VI), as confirmed by
XPS evidence [40]. Moreover, EXAFS analysis have also revealed that biogenic siderite
was able to carry out some reduction of U(VI) to U(IV) [39]. Thus one possible
interpretation of the results is the EC generated more reactive secondary minerals
contributing to the enhanced rate of U(VI) reduction. It should be noted that biogenic
189
minerals which tend to be more amorphous and less crystalline have been shown to be
more reactive than their more crystalline counterparts [39].
5.5.6. Evidence of anoxic corrosion of Fe0
Evidence that the corrosion was taking place under both abiotic and biotic
conditions was observed in this study through the significant release of soluble Fe2+ and
H2 gas. Approximately half of the added iron was recovered as Fe2+ during either the
abiotic or biotic pre-incubation of Fe0. During abiotic anoxic corrosion, H2 accumulated
in just a matter of days to yields of 72.4 and 54.3% of the maximum depending on
whether the maximum was defined by conversion to Fe2+ or to Fe3O4. The yield of H2
was significantly lower in the biotic incubations, which clearly indicates cathodic H2 was
being consumed by microorganisms. There is important literature evidence indicating
that cathodic H2 from Fe0 corrosion can be consumed by sulfate-reducing bacteria,
methanogens and autotrophic denitrifiers [19, 20, 60].
5.5.7. Biological Fe(III) reduction by cathodic H2 and its implication on U(VI) reduction
In this study, H2 consumption by the enrichment culture was shown to be
dependent on the presence of Fe(III) containing minerals, magnetite and ferrihydrite.
190
These observations provided evidence that H2 is being used to reduce Fe(III) minerals.
This is further supported by the fact that Fe2+ formation in this experiment was dependent
on the presence of inoculum and H2. It is known that certain sulfate-reducing bacteria –
such as Desulfovibrio desulfuricans – may utilize H2 as electron donor to reduce Fe(III)
oxides, with the formation of magnetite and siderite [61]. Also, some dissimilatory iron
reducing bacteria (DIRB) are able to reduce poorly crystalline and/or aqueous complexes
of Fe(III) with H2 [49, 62-66]. Specifically, Geobacter sulfurreducens can reduce Fe(III)
from ferrihydrite to form magnetite [66]. G. bremensis sp. nov. and G. pelophilus sp.
nov. have also demonstrated to use H2 for the reduction of ferrihydrite [67]. Some species
of methanogens demonstrated to utilize H2 to reduce Fe(III), including the species
Methanosarcina barkeri MS and Methanococcus voltaei A3 [68]. Similarly, some
thermophilic and hyperthermophilic bacteria can act towards soluble Fe(III) complexes in
the presence of H2 [69-72]. Although no other reports have been found on the use of H2
for the reduction of Fe(III) in magnetite, the results suggest that H2 is being consumed by
the enrichment culture for the reduction of Fe(III) in the iron containing minerals. Thus
the EC could contribute to generating more reduced secondary minerals by reducing
Fe(III) to Fe(II). Minerals completely composed of Fe(II) as the iron species such as a
siderite or vivianite are known to serve as a surface for the reduction of U(VI) [40].
191
5.5.8. Passivation of iron
The biggest difference between the abiotic and biotic anoxic corrosion was the
passivation of the Fe0. Biotic pre-incubation of Fe0 increased the reactivity of the iron
both in terms of Fe2+ released and U(VI) reduction. Abiotic pre-incubation of the Fe0, had
the opposite effect of rendering the iron non-reactive with U(VI) and preventing any
further release of Fe2+. These findings indicate a complete passivation had taken place.
The passivation was so severe that inoculating the abiotically pre-incubated Fe0 could not
overcome the passivation. If U(VI) was present during an abiotic incubation of Fe0, then
the complete passivation did not take place and the Fe0 continued to be reactive with
U(VI) over long incubation periods albeit that the rate was slower than in biotic
incubations. Taken as a whole, the results suggest that as long as either microorganisms
and/or U(VI) were reacting with the Fe0 surfaces complete passivation did not occur.
Passivation of Fe0 has been postulated to occur due to the formation of magnetite [73].
Thus the proven capacity of the enrichment culture to reduce magnetite may have the
benefit of refreshing the Fe0 surfaces or preventing maturation of Fe3O4 crystals that are
less reactive than biogenic Fe3O4 [39]. The net result could be a higher reactivity of Fe0
surfaces for U(VI) reduction in cultures by prevention of passivation. The extensive
release of Fe2+ by the enrichment culture would fit this hypothesis. Figure 5.18
summarizes the main hypotheses developed in this work.
192
Figure 5.18. Schematic of hypothesis of the role of microorganisms in the reduction of
Fe(III) from corrosion products. (1) Anoxic corrosion of Fe0 with formation of H2 and
magnetite (Fe3O4(s)); (2) biotic reduction of Fe(III) in magnetite (Fe3O4(s)) with H2 and
release of soluble Fe2+; (3) formation of Fe(II) secondary minerals, such as siderite
(FeCO3(s)), pyrite (FeS(s)) or vivianite (Fe3(PO4)2(s)), and (4) reduction of U(VI) to
insoluble U(IV) with (a) de
de-passivated Fe0 or (b) Fe(II) secondary minerals.
5.5.9.
.5.9. Bacterial population
In this study, the Dechloromonas phylotype showed a high 16S rRNA gene
sequence similarity
ity to several well
well-studied Dechloromonas species of (per)chlorate
reducing bacteria such as Dechloromonas sp. JM (98%) [74] and Perchlorate-reducing
Perchlorate
bacterium EAB2 (98%) [75]. Literature evidence shows that Dechloromonas has been
previously found adjacent to uranium
uranium-contaminated sites [76, 77],, although no uraniumuranium
reducing function has been reported from this ribotype. Some species of Dechloromonas
have been also reported for their use of Fe(II) as alternative ele
electron
ctron donor for the
193
reduction of chlorate [78] and nitrate [79] under anaerobic conditions. For example, the
Dechloromonas clones have 98% similarity to Dechloromonas agitata strain CKB, which
is able to oxidize the Fe(II) with (per)chlorate under strictly anaerobic conditions [80,
81].
The Stenotrophomonas phylotype observed in this study shows 99% 16S rRNA
gene sequence similarity to an uncultured bacterium clone R2-B12, which was identified
from an autrotrophic denitrifying granular sludge [82]. It has also been closely associated
to a culture capable of iron-oxidation and of growing on either FeS or FeCO3 [83]. On the
other hand, Stenotrophomonas has been reported for non-metabolic accumulation of
uranium polyphosphates in the periplasm [84].
The EC is clearly responsible for the Fe(III) reduction. Yet the phylotypes closely
related Dechloromonas and Stenotrophomonas bacteria are too date not known for that
function. Direct evidence is not available to fully understand what role these bacteria
have in the EC and if the identified bacteria are responsible for Fe2+ oxidation in the EC
its not clear what the electron-acceptor is for that reaction. Their presence may reflect a
highly preferential amplification by PCR, which could overshadow other bacteria in the
EC based on the clone library approach.
194
5.5. Conclusions
Enrichment cultures, which accelerated U(VI)-reduction with Fe0 compared to the
abiotic rate, were successfully maintained for more than 20 transfers. Bacteria in the
enrichment culture can modify Fe0 during anoxic corrosion making it more reactive with
U either due to passivation prevention and/or due to formation of reactive secondary
minerals. Precise mechanisms of U(VI) reduction with corrosion products need to be
studied further. Cathodic H2 from Fe0 corrosion was consumed by one or more members
of the enrichment-culture to support Fe(III) reduction in magnetite and ferrihydrite
minerals. A clone library revealed the predominant OCR amplified clones to be closely
related to Dechloromonas and Stenotrophomonas, both are known from U contaminated
sites and the former is known for anoxic Fe2+ oxidation.
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208
CHAPTER 6
TOXICITY OF URANIUM TO MICROBIAL PROCESSES IN ANAEROBIC
SLUDGE
6.1. Abstract
Contamination with high levels of uranium and nitrate is a common problem in
groundwater impacted by uranium mine tailings. Bioremediation approaches have shown
promise in the remediation of uranium and nitrate.
Nevertheless, there is limited
information on the inhibitory impact of uranium(VI) on the microorganisms occurring in
uranium bioremediation sites. The objective of the present work is to determine the
inhibitory potential of uranium to different microbial populations, including
methanogenic, denitrifying and uranium-reducing microorganisms, which are commonly
present in these sites. For this purpose, batch toxicity assays were conducted with a
sulfur-oxidizing denitrifying mixed culture and a methanogenic consortium. Results
indicated a very distinct level of toxicity depending on the targeted microbial population.
U(VI) caused severe inhibition of acetoclastic methanogenesis, with a 50% inhibiting
concentration (IC50) of only 0.16 mM. Denitrifying populations were also impacted by
the presence of uranium, but their sensitivity depended on the electron donor utilized.
Sulfur oxidizing denitrifiers were the least affected by U(VI) (IC50 = 0.32 mM), followed
209
by H2- and acetate-utilizing denitrifiers (IC50 of 0.20 and 0.15 mM, respectively). In
contrast to the inhibitory effect of U(VI) towards methanogenesis and denitrification,
exposure to U(VI) concentrations up to 1.0 mM did not inhibit the rate of U(VI)
reduction by the methanogenic- and the denitrifying mixed cultures with H2 as electron
donor. On the contrary, a considerable increase in the uranium reducing activity of both
cultures was observed with increasing uranium concentrations. This seems to suggest
that denitrification, methanogenesis and uranium reduction are carried out by different
microorganisms. Results from this study provide insights on the potential toxicity of
U(VI) over microbial processes common in bioremediation systems for uranium and
nitrate.
6.2. Introduction
Large remediation efforts are being undertaken to eliminate uranium from
groundwater exposed to adjacent uranium mining and milling sites. Uranium-bearing
groundwater results from the inadequate confinement of mill tailing wastes in
repositories that may leach or permeate [1] as well as from the high natural levels of this
radionuclide occurring in certain zones [2]. Very often, uranium-impacted water is also
contaminated with high levels of nitrate originating from nitric acid used in uranium
processing [3].
210
From all the uranium forms found in the environment, tetravalent uranium
(U(IV)), which is often present as the mineral uraninite (UO2), is the most stable and
insoluble; whereas soluble hexavalent uranium (U(VI)), generally occurring as uranyl ion
(UO22+), is the most reactive species [1]. Hexavalent uranium is toxic since it can disrupt
the normal functions of kidneys [4]. Due to its human toxicity, the presence of this form
of uranium in drinking water is of concern.
As a result, the U.S. Environmental
Protection Agency has set a maximum contaminant level for uranium in drinking water
of 30 µg/L [4].
Different methods are being applied to attain remediation of U(VI) contaminated
water, including physical-chemical methods such as anion exchange [5], and coagulation
with iron or aluminum salts [6], as well as biological methods such as biosorption [3, 7],
bioaccumulation [8], biomineralization [9], and bioreduction [10, 11].
Microbial
immobilization of uranium by anoxic bioreduction of U(VI) to U(IV) is one of the most
attractive and effective methods for the remediation of U(VI) contaminated water [10].
A number of anaerobic bacteria have demonstrated capacity to perform this
process in the presence of different electron donors, such as ethanol and acetate [11-13].
The majority of the microorganisms known to be involved in uranium remediation
processes are anaerobes that belong to the group of sulfate-reducing bacteria, such as
Desulfovibrio spp. [14-17], as well as to the Fe(III)-reducing bacteria, including some
species of the genera Shewanella [18-22] and Geobacter [23-26]. Other species have
211
also been reported, such as Clostridium spp. [27, 28]. In conditions where nitrate (NO3-)
is present, such as in mill tailings, Thiobacillus denitrificans is known to be able to link
the oxidation of U(IV) to the reduction of NO3- [29].
Microorganisms known for their ability to reduce U(VI) have been found in
anaerobic methanogenic biofilms from full-scale up-flow anaerobic sludge blanket
(UASB) reactors treating brewery and distillery wastewater (Eerbeek sludge) [30-33]. In
fact, recent studies have confirmed that methanogenic Eerbeek biofilms have an intrinsic
activity towards U(VI) reduction to U(IV) [34], even under continuous operation
conditions with high U(VI) loading rates [35].
Furthermore, integrated biological
systems for the sequential treatment of nitrate and U(VI) in groundwater have been
developed with successful results [36].
The potential toxicity of uranium towards the microbial communities present in
denitrifying and U(VI) reducing systems used to remediate groundwater contaminated by
mining activities is poorly understood. Studies published to date have only considered a
few of these bacterial species, including Pseudomonas aeruginosa, Clostridium sp.
ATCC 53464 and Thermoterrabacterium ferrireducens, and have confirmed that U(VI)
can inhibit microbial growth at concentrations greater than 1,000, 50 and 1,190 mg L-1,
respectively [27, 37, 38]. The toxicity of soluble U(VI) to bacteria has been attributed to
the blockage of some protein sites which are essential for their binding to DNA, which
interferes with gene expression and DNA repair [39].
212
In the present work, the possible inhibitory effects of U(VI) to diverse groups of
microorganisms commonly present in anoxic or anaerobic bioremediation systems were
assessed in order to determine U(VI) threshold levels to prevent microbial inhibition and
facilitate bioremediation efforts.
6.3. Materials and methods
6.3.1. Biomass sources
A thiosulfate-adapted mixed culture obtained from a laboratory-scale reactor
(10.6% VSS) was utilized in the autotrophic denitrification toxicity assays using
elemental sulfur as electron donating substrate [40]. Anaerobic granular biofilms
acquired from an upflow anaerobic sludge blanket (UASB) reactor treating recycled
paper wastewater (Eerbeek, The Netherlands) (11.3% VSS) were used as inoculum for
toxicity assays over uranium reduction, methanogenesis, and denitrification with H2 and
acetate. Sludge biofilms were stored anaerobically at 4ºC and previous to their use in
toxicity assays, they were washed and sieved as described earlier [34].
213
6.3.2. Basal medium
The mineral basal medium used in uranium-reduction toxicity assays (M1)
consisted of (mg L-1): NH4HCO3 (5.0), K2HPO4 (2.0), MgSO4·7H2O (2.5), Ca(OH)2
(1.0), yeast extract (0.33), as well as the following concentration of trace elements (in µg
L-1): H3BO3 (0.5), FeSO4·7H2O (28.0), ZnSO4·7H2O (1.06), MnSO4·H2O (4.15),
(NH4)6Mo7O24·4H2O (2.0), AlK(SO4)2·12H2O (1.75), NiSO4·6H2O (1.13), CoSO4·7H2O
(23.6), Na2SeO3·5H2O (1.0), Na2WO4·H2O (5.0), CuSO4·5H2O (1.57), EDTA (10.0), and
resazurin (2.0). After adjusting to a pH value of 7.0, the media was amended with
NaHCO3 to a final concentration of 5.0 g L-1 (59.0 mM). The composition of the basal
mineral medium for methanogenic toxicity assays (M2) was the same, with only a
modification in the concentration of NaHCO3 to 2.5 g L-1 (29.8 mM), which was added
after adjusting the pH to 7.5. Finally, the mineral medium used in the denitrification
toxicity studies (M3) was the following (mg L-1): NH4HCO3 (5.0), K2HPO4 (2.0),
KH2PO4 (0.5), MgSO4·7H2O (3.0), with the same final concentration of trace elements
than for M1. The medium was adjusted to a pH value of 7.0 and subsequently amended
with NaHCO3 to a final concentration of 2.0 g L-1 (23.5 mM).
214
6.3.3. Batch toxicity bioassays
Table 6.1 summarizes the conditions evaluated in the different toxicity
experiments. Experiments were all carried out in 160-mL serum flasks (Wheaton,
Millville, NJ, USA).
215
Table 6.1. Summary of experimental conditions applied in the toxicity assays.
Toxicity Assay
Sludge
Inoculum
conc.
[g VSS/L]
Uranium reduction
Eerbeek
0.5
M1
H2
19.1
0 – 1.0
H2/CO2 c
U
Methanogenesis
Eerbeek
1.5
M2
Acetate
31.3b
0 – 2.1
N2/CO2 d
CH4
Denitrification
Thiosulfate
adapted sludgea
0.5
M3
Sulfur
15.0
0 – 0.6
N2/CO2 d
NO3-/NO2-, U
Eerbeek
0.5
M3
Acetate
4.0
0 – 0.6
N2/CO2 d
NO3-/NO2-, U
Eerbeek
0.5
M3
H2
25.0
0 – 0.6
H2/CO2 c
NO3-/NO2-, U
a
b
c
d
Basal
medium
e-donor
e-donor conc.
[mmol Lliquid-1]
U(VI) conc.
[mM]
Headspace
content
Analyte
From a laboratory-scale denitrifying reactor, as described in section 5.3.1.
Concentration corresponds to 2 g COD L-1.
H2/CO2 gas mixture in a proportion 80:20 (v/v).
N2/CO2 gas mixture in a proportion 80:20 (v/v).
216
6.3.3.1. Methanogenic toxicity bioassays
The toxicity of U(VI) to methanogens was evaluated in serum flasks containing
25 mL of basal media (M2) and 135 mL of headspace. Assays were inoculated with
Eerbeek sludge to a final concentration of 1.5 g VSS L-1 and supplied with acetate or
hydrogen as electron donor (e-donor). Acetate (2.56 g L-1, equivalent to 2 g COD L-1)
was supplied using a stock solution containing 64.0 g L-1 sodium acetate (NaC2H3O2).
The flask headspace was flushed with N2/CO2 (80:20, v/v) as described above to ensure
anaerobic conditions. In treatments with H2 as e-donor, the sealed bottles were first
flushed with N2/CO2 (80:20, v/v) and then with an overpressure of 0.8 atm of a gas
mixture of H2/CO2 (80:20, v/v). Treatments without uranium were used as uninhibited
controls.
Bottles designated for treatments and controls were all performed in
quadruplicate, and pre-incubated overnight in the dark at 30±2°C in an orbital shaker
(150 rpm) to activate methanogenic microorganisms. The next day, the headspace was
flushed again with N2/CO2 to purge of all traces of CH4 formed. Then, U(VI) was
provided from a 10 mM stock solution of UO2Cl2 · 3H2O, and bottles were incubated at
30 ºC. Samples of the headspace were obtained periodically to monitor the production of
methane.
217
6.3.3.2. Denitrification toxicity bioassays
The toxicity of uranium towards nitrate reducing microorganisms in two different
mixed cultures (Eerbeek methanogenic sludge and thiosulfate-adapted sludge) was
assayed in 160-mL serum flasks containing 50 mL of basal medium (M3) supplied with 4
mM of NO3- (0.25 g L-1) and known amounts of the respective electron donor (elemental
sulfur or S0, hydrogen or acetate) as is indicated in Table 6.1. All bottles were inoculated
with 0.5 g VSS L-1 of the respective sludge and spiked with varying volumes of a 10 mM
stock solution of UO2Cl2 3H2O to attain uranium concentrations ranging from 0.005 to
0.6 mM. Finally, the liquid was flushed with N2:CO2 (80:20, v/v) to create anaerobic
environment (as described before), the flasks were sealed and then incubated at 30±2°C.
Assays without uranium were used as uninhibited controls.
Samples were taken
periodically for analysis of soluble U, nitrate (NO3-) and nitrite (NO2-).
6.3.3.3. Toxicity over uranium-reduction activity
Experiments to monitor the inhibitory effect of increasing concentrations of
uranium over the U(VI) reduction rate consisted of 100 mL of basal media (M1) and 60
mL of headspace. All bottles were supplied with the desired concentrations of granular
sludge (Table 6.1). Hexavalent uranium was provided in the form of uranyl chloride
trihydrate (UO2Cl2 · 3H2O) by taking different aliquots from a 10 mM U(VI) stock
218
solution to attain the desired range of concentrations indicated in Table 6.1. Assays
without uranium were run in parallel to serve as uninhibited controls.
To ensure
anaerobic conditions, the headspace of the flasks was flushed with N2/CO2 gas (80:20,
v/v) for 30 s. The flasks were sealed with butyl rubber septa and aluminum crimp seals,
and the headspace was flushed again by continuous flow for 4 min through inlet and
outlet needles inserted in the septa of the bottles [34]. Afterwards, the headspace was
supplied with a gas mixture of H2/CO2 (80:20, v/v) at an overpressure of 0.8 atm. All
treatments and controls were performed in duplicated replicates, and incubated in the
dark at 30±2°C in an orbital shaker at 150 rpm. Liquid samples were collected
periodically to monitor the soluble uranium concentration.
The maximum specific uranium reducing activity (mg U(VI)-removed g VSS-1 d1
), denitrifying (mg NO3¯-removed g VSS-1 d-1) and methanogenic (mg CH4-COD-
formed g VSS-1 d-1) activities were calculated from the slope of the U(VI) removed,
nitrate concentration and cumulative methane production; respectively, and biomass
concentration versus time (d), as the mean value of quadruplicate or duplicate assays. In
each case, the maximum specific activity at a given U(VI) concentration was determined
during the time period when the uranium-free control displayed maximum specific
activity. The inhibition observed in the denitrification- and methanogenic bioassays was
calculated as shown below:
219

Maximum ⋅ Specific ⋅ Activity ⋅ at ⋅ the ⋅Tested ⋅ Concentration 
Inhibition ⋅ (%) = 100 − 100 ⋅

Maximum ⋅ Specific ⋅ Activity ⋅ of ⋅ the ⋅ Control


The initial concentrations of U(VI) causing 20, 50 and 80% reduction in activity
compared to an uninhibited control were referred to as IC20, IC50 and IC80, respectively.
These values were calculated by interpolation in the graph plotting the inhibition
observed (expressed as percent) as a function of the inhibitor concentration. Unless
otherwise indicated, reported inhibitory concentrations are average values of
quadruplicate or duplicate assays and corresponding standard deviations.
6.3.4. Analytical techniques
Samples for analysis of soluble uranium were pipetted into Eppendorf
TM
centrifuge 1.5-mL tubes and centrifuged at 10,000 rpm (RCF of 10,621 x g) for 10 min.
Immediately afterwards, supernatant was separated from the solid pellet and transferred
to a 3% HNO3 solution. Soluble uranium was measured by using an Inductively Coupled
Plasma – Optical Emission Spectrometry (ICP-OES) system (Optima 2100 DV from
Perkin-Elmer, Shelton, CT, USA) at a wavelength of 385.958 nm. The detection limit for
U was 10 µg L-1.
220
Nitrate and nitrite were measured by ion chromatography (IC) with suppressed
conductivity detection using a DIONEX 500 IC-3000 system fitted with a Dionex IonPac
AS16 analytical column (4 mm x 250 mm) and a AG16 guard column (4 mm x 40 mm),
and a CD 20 conductivity detector (Dionex, Sunnydale, CA, USA). The mobile phase
was 5 mM KOH from 0 to 6 min, 5 to 30 mM KOH from 6 to 10 min and 30 mM KOH
from 10 to 15 min. The injection volume was 75 µL. The detection limit for nitrate and
nitrite was 0.01 mg L-1.
Methane content in gas samples was analyzed by gas chromatography (GC) using
a Hewlett-Packard 5890 Series II system (Agilent Technologies, Palo Alto, CA, USA),
equipped with a flame ionization detector and a Stabilwax-DA fused silica capillary
column (30 m length × 0.53 mm ID, Restek Corporation, Bellefonte, PA, USA). The
injection volume was 100 µL. Helium was used as the carrier gas at a flow rate of 18 mL
min-1 and a split flow of 85 mL min-1. The temperatures of the oven, injector port and
detector were 140, 180 and 250ºC, respectively.
Volatile suspended solids (VSS) and total suspended solids (TSS) were evaluated
according to Standard Methods [41].
221
6.3.5. Chemicals
Ammonium bicarbonate (NH4HCO3, 21.3-21.7% as NH4+) and nitric acid (HNO3,
70%) were purchased from Fisher Chemical (Fair Lawn, NJ, USA). Magnesium sulfate
(MgSO4·7H2O), calcium hydroxide (Ca(OH)2), potassium phosphate dibasic (K2HPO4,
>99.0%) and sodium acetate anhydrous (NaC2H3O2, >99.0%) were all obtained from J.T.
Baker (Phillipsburg, NJ, USA). Potassium nitrate (KNO3, 99.0%) and sublimed powder
sulfur (S0, >99.0%) were purchased from EMD Chemicals (Gibbstown, NJ, USA). Yeast
extract was purchased from BD (Sparks, MD). Sodium bicarbonate was obtained from
Pfaltz & Bauer (Waterbury, CT, USA). Finally, uranyl chloride trihydrate (UO2Cl2 ·
3H2O) was obtained from International Bio-Analytical Industries, Inc. (Boca Raton, FL,
USA).
6.4. Results and discussion
6.4.1. Methanogenic toxicity
Batch bioassays were conducted with a mixed anaerobic consortium (Eerbeek) to
investigate the toxic effects of U(VI) to acetate-utilizing methanogens. Figure 6.1A
shows the time course of methane (CH4) formation during this experiment. Figure 6.1B
222
illustrates the normalized acetoclastic methanogenic activity of Eerbeek sludge as a
function of the initial U(VI) concentration. Both figures indicate a steep decrease in the
rate of CH4 production with increasing U(VI) concentrations, reaching an almost
complete inhibition at 2.1 mM (8.7% activity of the uninhibited control) (Figure 6.1B).
223
Figure 6.1. Toxic effect of increasing uranium(VI) concentrations over the acetoclastic
methanogenic activity of a mixed microbial culture ((Eerbeek sludge). (A) Time course of
CH4 concentration (%). Concentration of U(VI) in mM: (—○—), 0; (—
—∆—), 0.05; (—
□—), 0.2; (—▲—) 0.4;; ((——), 0.6; (——), 1.0; (—●—), 2.1. (B) Methanogenic
activity with respect to the initial concentration of U(VI).
224
Table 6.2 summarizes the IC20, IC50 and IC80 values determined for U(VI) in the
acetoclastic methanogenic toxicity bioassays. The low inhibitory concentrations
determined (IC50 = 0.16 mM) indicate that the methanogenic consortium present in the
Eerbeek sludge was particularly sensitive to U(VI). To the best of our knowledge, this is
the first published study that evaluates the inhibitory effects of U(VI) towards
methanogens.
Methanogens are expected to be important players in the microbial
communities found in engineered anaerobic systems for uranium bioremediation [42-44].
Table 6.2. Concentrations of uranium causing 20% (IC20), 50% (IC50) and 80% (IC80)
inhibition of the activity of methanogenic and denitrifying microorganisms present in the
biomass sources tested.
Uranium concentration (mM)
Microbial Activity
Substrate
IC20
IC50
IC80
Methanogenesis
Acetate*
0.030
0.160
0.880
Denitrification
Acetate*
0.054
0.148
0.250
0.100
0.200
0.380
0.100
0.320
>0.600
H2
*
Sulfur (S0)**
* Eerbeek sludge
** Thiosulfate-adapted sludge
225
6.4.2. Toxicity to denitrifying microorganisms
6.4.2.1. Denitrification with elemental sulfur
The inhibitory effect of U(VI) towards microorganisms linking denitrification to
elemental sulfur (S0) oxidation was investigated in experiments using a thiosulfateadapted mixed culture. Figures 6.2A and 6.2B depict the time course for nitrate (NO3-)
and nitrite (NO2-) in assays amended with initial U(VI) concentrations ranging from 0 to
0.60 mM. A considerable decrease in the rate of NO3- reduction (panel 6.2A) was
observed with increasing U(VI) concentration. The rate of NO2- reduction (panel 6.2B)
was also affected by the presence of U(VI).
An initial accumulation of NO2- was
observed which peaked at day 4-6, depending on the assay, and was consistent with the
stoichiometry of the conversion from NO3- to NO2-. NO2- consumption started as soon as
NO3- was depleted. The impact of increasing U(VI) concentrations on the normalized
NO3- and NO2- reducing activities of the mixed culture is illustrated in Figure 6.3A.
These results confirm that both NO3- and NO2- reducing microorganisms are inhibited by
U(VI), with nitrite-reducers showing a somewhat higher sensitivity to this toxicant.
226
Figure 6.2. Effect of uranium
uranium(VI) concentration on nitrate reduction (A) and nitrite
reduction (B) by a thiosulfate
thiosulfate-adapted mixed culture utilizing S0 as electron donor.
donor
Concentration of U(VI) in mM: ((—○—), 0; (—□—), 0.005; (—∆—),, 0.02; (—▲—)
(
0.1; ((—●—), 0.2; (——), 0.4; (——), 0.6.
227
Figure 6.3. Role of initial uranium(VI) concentration on the normalized denitrification
activity (respect to the uninhibited control) of (A) a thiosulfate-adapted
adapted inoculum using
S0 as electron donor;; (B) an anaerobic mixed culture (Eerbeek sludge) utilizing acetate as
electron donor,, and (C) an anaerobic mixed culture ((Eerbeek sludge) utilizing H2 as
electron donor. Legends: (—●—), NO3- reducing activity; (—○—), NO2- reducing
activity.
228
6.4.2.2. Denitrification with acetate and H2 as electron donors
The inhibitory effect of U(VI) towards denitrifying organisms utilizing acetate
and H2 as energy source was investigated in assays inoculated with an anaerobic mixed
culture (Eerbeek sludge). Although this inoculum was obtained from a methanogenic
reactor, the mixed culture also displayed sizable denitrifying activity with the electron
donors utilized.
Figure 6.4A shows the evolution of the NO3- concentration over time at
concentrations of U(VI) ranging from 0 to 0.60 mM in assays utilizing acetate as electron
donor. The figure shows that concentrations of U(VI) exceeding 0.10 mM led to a
decrease in nitrate reducing activity. In this case, no significant accumulation of NO2was observed over the duration of the experiment. Figure 6.3B shows the normalized
denitrification activities as a function of increasing U(VI) concentration. The IC50 value
of U(VI) towards acetate-utilizing denitrifiers was 0.15 mM U(VI) and complete
inhibition was observed at a concentration of 0.60 mM (Table 6.2).
229
Figure 6.4. Effect of uranium
uranium(VI) concentration on nitrate reduction by an anaerobic
mixed culture (Eerbeek
Eerbeek sludge) utilizing (A) acetate as electron donor,, and (B) H2 as
electron donor. Concentration of U(VI) (in mM): (—○—), 0; (—∆—),, 0.02; (—▲—)
(
0.1; ((—●—), 0.2; (——), 0.4; (——), 0.6.
230
Figure 6.4B shows the time course of NO3- in a similar experiment with H2 as
electron donor and confirms the inhibitory impact of U(VI) on the autotrophic H2utilizing denitrifying population. A very low accumulation of NO2- was detected initially
but it decreased to negligible levels after day 4 (results not shown). Figure 6.3C displays
the normalized denitrifying activities determined in this experiment. The IC50 value of
U(VI) was 0.20 mM, and complete inhibition was observed at concentrations exceeding
0.38 mM U(VI) (Table 6.2). U(VI) was more inhibitory towards acetate- and H2-utilizing
denitrifiers compared to sulfoxidizing denitrifiers as indicated by the higher inhibitory
concentrations determined in the latter case (Table 6.2, Figure 6.3).
Published data on the toxic effects of U(VI) towards denitrifying bacteria are very
scarce. We are only aware of the work conducted by Bencheikh-Latmani and coworkers
[45] reporting complete inhibition of citrate degradation with nitrate as electron acceptor
by Pseudomonas fluorescens, a known denitrifying bacteria, at U(VI) concentrations
above 5 µM. The authors attributed the inhibition observed to U(VI) biosorption on the
cell envelop of these microorganisms, a phenomenon observed both with living and dead
cells. The inhibitory concentrations observed in our study (Table 6.2) are many fold
higher compared to those reported by Bencheikh-Latmani and coworkers. The higher
tolerance exhibited by denitrifiers in our assays may be related to the use of thick
sulfoxidizing- and anaerobic biofilms which afford some protection against exposure of
the denitrifiers to the toxicant.
The discrepancy may also be related to the higher
231
sensitivity of Ps. fluorescens to U(VI) compared to the denitrifying bacteria in the inocula
utilized in the current study.
Although the mechanisms by which uranium in the form of uranyl ion (UO22+)
inhibits microorganisms are not clearly understood, U(VI) is believed to interfere with
the function of DNA-binding proteins, resulting in disruption of gene expression [39].
More recently, it has been reported that the uranyl ion may hinder the binding of Ca2+ and
other cations to proteins [46]. A study by Tuovinen and coworkers on the effects of
U(VI) over the chemolithoautotroph Thiobacillus ferrooxidans [47, 48] revealed that
uranium can disrupt CO2 fixation and ferrous oxidation activity, but that EDTA or other
competing cations can reduce the inhibitory effects of U(VI).
6.4.3. Inhibition of Uranium Reduction Activity
The inhibitory potential of hexavalent uranium to U(VI)-reducing activity of
microorganisms in an anaerobic mixed culture (Eerbeek sludge) was evaluated in shaken
batch biossays. This unadapted sludge was previously shown to have a high intrinsic
capacity to reduce U(VI) to U(IV) utilizing H2 as electron donor [34]. Figure 6.5 shows
the time course for U(VI) reduction by Eerbeek sludge spiked with initial U(VI)
concentrations ranging from 0 to 1.0 mM in assays utilizing H2 as electron donor. Within
the range of U(VI) tested, a high activity towards U(VI) reduction was observed, without
232
any observable lag phase.
In fact, the rate of uranium reduction increased as the
concentration of U(VI) increased from 0 to 1.0 mM (Figure 6.6).
Figure 6.5. Effect of increasing uranium
uranium(VI)
I) concentrations over the U(VI)-reducing
U(VI)
activity of an anaerobic mixed culture ((Eerbeek sludge). Concentration of U(VI) in mM:
(—□—), 0.03; (—
—▲—) 0.2; (——), 0.4; (——), 0.8; (—●—
—), 1.0.
233
Figure 6.6. Impact of increasing initial uranium(
uranium(VI)
VI) concentrations on the rate of
uranium removal by Eerbeek sludge in assays supplied with H2 (—●—
—) and acetate
(—○—).
Although the impact of U(
U(VI)
VI) on the uranium reducing activity of methanogenic
enrichments has not been reported previously, several previous studies have considered
the effect of uranium on U(VI) bioreduction. U(VI) has been shown to inhibit uranium
reducing activity and/or growt
growth of U-reducing
reducing microorganisms, including purepure and
mixed cultures, at low to intermediate concentrations. Sani et al. [49] reported complete
inhibition of the growth and U(VI) reducing activity of the sulfate reducing bacterium,
bacteriu
Desulfovibrio desulfuricans G20, at U(VI) concentrations above 0.14 mM. Likewise,
Nyman et al. [50] observed complete growth inhibition of U(VI) reducingreducing and sulfate
234
reducing enrichment cultures obtained from a remediation site where biostimulation was
attempted through ethanol addition at U(VI) exceeding 1.60 mM. In the latter study, the
IC50 value determined for U(VI) was around 0.13 mM. A considerable higher tolerance
to U(VI) was observed for the anaerobic inoculum used in this study which suggests the
presence of microbial communities displaying relatively high resistance to this chemical.
Interestingly, Nyman and coworkers [50] have shown that sequences relative to
Clostridia spp. microorganisms, which are U(VI) reducing bacteria previously detected in
the Eerbeek inoculum [30-32], are relatively resistant to U(VI). Microorganisms such as
Thermoterrabacterium ferrireducens have also been reported to be relatively tolerant to
U(VI) and levels of this species exceeding 2.5 mM were required to cause inhibition of
uranium reducing activity and cell growth [38].
A gradual decrease in the concentration of soluble uranium was also observed in
the acetoclastic methanogenic assays with Eerbeek sludge during the course of the
experiment (results not shown), suggesting that the removal was due to microbial
reduction rather than chemical precipitation. Although acetate is not a known electron
donor of uranium-reducing bacteria, excess endogenous substrate in the anaerobic
inoculum utilized has been shown earlier to support uranium reduction for extended time
periods of up to 13 months [35]. Figure 6.6 compares the removal rates of U(VI)
observed in the acetoclastic methanogenic assays to those determined previously with H2
as electron donor. In contrast to the assays using H2 which showed enhanced U-reducing
activities with U(VI) concentrations up to 1.0 mM, the rates of U(VI) removal in the
235
acetoclastic methanogenic experiments were inhibited at concentrations exceeding 0.6
mM.
This could be due to the lower U(VI) reducing potential of the endogenous
substrate in the sludge compared to H2 [34].
The U(VI) level in denitrifying treatments with S0 and acetate as electron donors
remained relatively constant after 2 days of incubation, and only a small initial loss was
observed in some treatments. As an example, in denitrifying assays with acetate as
electron donor, 10.4 to 15.9% of the U(VI) added was removed after 2 days in the
treatments spiked with 0.10 mM or higher concentrations of uranium, and the losses were
negligible in assays with lower U(VI) levels. These trends suggest that the observed
U(VI) elimination was due to partial precipitation of uranium with some components of
the media and not to microbial reduction. It has been observed that ligands such as
phosphate can precipitate soluble U(VI) [51]. In contrast with treatments amended with
S0 and acetate, a steady decrease in the concentration of U(VI) was observed during the
course of the experiment, suggesting the ability of microorganisms in the inoculum to
reduce U(VI) using H2 as electron donor (results not shown). The rate of U(VI) removal
increased from 8.3 to 83.8 µM/day when the initial U(VI) concentration was raised from
0.02 to 0.60 mM, indicating that this species did not inhibit the U(VI) reducing activity
of the inoculum. These findings are in contrast with the severe inhibition exerted by
U(VI) towards the denitrifying activity of microorganisms in the same microbial culture
and suggest that the microorganisms responsible for U(VI) reduction and for
denitrification might not the same.
236
6.5. Conclusions
The present communication addresses the possible effects of U(VI) over the
communities present in anaerobic granular sludge considered for application in uranium
and nitrate bioremediation. Uranium and nitrate are two contaminants which often cooccur in groundwater impacted by uranium mill tailings. Results from this study confirm
that a mixed culture biofilm obtained from a full-scale methanogenic bioreactor displayed
high U(VI) reducing activity when supplied with H2 in the presence or absence of nitrate.
High U(VI) reducing activities were also observed in cultures spiked with acetate, an
organic acid which is not a known electron donor of uranium reducers.
Uranium
reduction in the latter assays was supported by endogenous substrate in the biofilms. In
contrast, U(VI) reduction was not observed under denitrifying conditions when acetate or
elemental sulfur were the only electron donors.
High concentrations of U(VI) (up to 0.6-1.0 mM, depending on the assay) did not
inhibit the U(VI) reducing activity of the mixed cultures utilized in this study.
Hexavalent uranium, however, was highly inhibitory to the methanogenic activity as well
as the nitrate- and nitrite reducing activity of the mixed cultures utilized as shown by IC50
values ranging from 0.15 to 0.32 mM.
These findings suggest that the microbial
population involved in U(VI) reduction might be different from those responsible for
methanogenesis and denitrification processes studied in this work. The maximum
concentrations of uranium for groundwater in US mill tailing contaminated sites often
237
vary from 7-42 µM (5.7 - 10 mg L-1) [52]; which are values considerably lower than the
inhibitory concentrations determined in this work. Taken as a whole, the results obtained
can be utilized to optimize the operation of bioremediation systems for uranium and
nitrate. In addition, they will be useful to facilitate understanding of microbial population
dynamics in systems where high, potentially inhibitory levels of uranium are present.
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246
CONCLUSIONS
The presence of hexavalent uranium (U(VI)) in groundwater represents a serious
public health concern due to its chemical toxicity. This research explored reductive
precipitation of soluble U(VI) to insoluble tetravalent uranium (U(IV)) as a low cost
bioremediation alternative for the treatment of uranium contaminated groundwater. Two
major approaches were investigated. The first was the applicability of anaerobic sludge
from wastewater treatment for U(VI) reduction. The second was to determine if
microorganisms can stimulate the chemical reduction of U(VI) by zero valent iron.
Methanogenic granular sludge from different upflow anaerobic sludge blanket
(UASB) reactors treating different industrial wastewaters were tested for their capacity to
support the reductive precipitation of U(VI). The anaerobic granules had an intrinsic
capacity to reduce U(VI), which is attributed to the large biodiversity existing within the
sludge. In addition, endogenous substrates originating from the decay of the sludge
provided a large pool of electron-equivalents for effective bioreduction in the absence of
exogenous electron donors.
The addition of H2 as an electron donor increased the
reduction rate of U(VI) to varying degrees depending on the level of endogenous
substrate present in the different sludge samples tested. H2 had the largest impact with
sludge samples characterized as having low levels of endogenous substrate. A sequential
247
extraction procedure was utilized to characterize the uranium after bioreduction with
sludge. The procedure which distinguishes between water soluble U(VI), U(VI) adsorbed
and insoluble U(IV) fractions, indicated that most of the uranium was present as U(IV).
Reoxidation of the immobilized uranium with O2 demonstrated that the mechanism of
U(VI) removal from solution was due to the reductive precipitation. X-ray diffraction
(XRD) further confirmed that the final form of uranium present was the insoluble
mineral, uraninite (UO2). The results taken as a whole suggested that the methanogenic
granules could be a very effective for application in bioremediation systems for treating
uranium.
The long-term application of methanogenic granules for uranium treatment was
investigated. Upward-flow continuous columns packed with sand and methanogenic
granular sludge were carried out in the presence and absence of ethanol as exogenous
electron donor. The studies demonstrated that reductive immobilization of U(VI) was
effectively sustained throughout the course of one year at high U(VI) loading rates (up to
246 µmol U(VI) Lr-1 d-1), with removal efficiencies as high as 99.8%. Sequential
extraction applied over the solids from the columns at the end of the continuous
experiment revealed that most of the U removed was recovered at the base of the
columns as U(IV), indicating that the main mechanism of immobilization was reductive
precipitation U(VI) to U(IV). The addition of ethanol as electron donor only had a shortterm contribution to improving the reductive activity in the column. The endogenous
substrate from sludge biomass decay was effective in sustaining reducing equivalents for
248
U(VI) reduction over the long term. Since nitric acid is used in mining for the extraction
of uranium from the ores, nitrate is a common co-occurring contaminant in uranium sites.
Therefore, nitrate needs to be taken into account in any bioremediation system with
uranium. Nitrate co-contamination was linked to the oxidative remobilization of the
biogenic U(IV) previously accumulated in the column. In addition, it was observed that
U(VI) had an inhibitory effect on denitrification. These results suggested that a prior
nitrate treatment stage is needed in any bioremediation system relying on anaerobic
sludge for uranium bioremediation.
Several remediation systems rely on zero-valent iron (Fe0) for the removal of
U(VI) from groundwater in contaminated sites. Fe0 is a very effective reducing agent that
can carry out the abiotic reductive precipitation from U(VI) to U(IV). A study was
conducted with the objective to demonstrate that microorganisms from an enrichment
culture (EC) developed from an Fe0-containing packed bed bioreactor were able to
enhance U(VI) reduction with Fe0. During the course of 28 months, the serial transfer of
EC could be sustained with an effective enhanced uranium reduction. It was shown that
the rate of biological U(VI) reduction with Fe0 was consistently from 3- to 27.4-fold
higher than that of the abiotic rate with Fe0 alone. A clone library of the EC indicated the
predominant microorganisms were from two bacterial ribotypes (closely related to
Stenotrophomonas and Dechloromonas). Characterization of the immobilized uranium by
sequential extraction and X-ray photoelectron spectroscopy (XPS) indicated the solid
phase speciation of U was predominantly U(IV). The possible roles of the EC on the
249
corrosion of Fe0 were evaluated in studies in which the Fe0 was preincubated with the
EC. The EC preincubated Fe0 had a higher U(VI) reduction rate a than treatments with
fresh Fe0, suggesting that bacteria may be preventing the passivation of the Fe0 surfaces
and/or generating biogenic secondary minerals that are more reactive with U(VI).
Abiotic corrosion of Fe0 was extensive with high yields of cathodic H2. During biological
corrosion of Fe0, consumption of cathodic H2 occurred, probably by some members of
the EC. Additional experiments revealed that the EC was consuming H2 for the reduction
of Fe(III)- corrosion mineral phases (such as magnetite), by forming Fe2+. Fe2+ could be
precipitated later as secondary Fe(II)-minerals with components of the media (such as
carbonate and phosphate). These secondary minerals may serve as chemical reductants of
U(VI). In this way, bacteria would be avoiding the passivation of Fe0 surfaces by
magnetite through the continuous consumption of the Fe(III) as observed from the large
Fe2+ release from Fe0 by the EC. The application of this work to the remediation field
would provide several benefits, including the decrease in Fe0 needed due to the microbial
U(VI)-reduction.
The potential toxicity of U(VI) to different microbial processes in anaerobic
sludge was evaluated. The processes evaluated with implications in the bioremediation of
uranium and nitrate, were methanogenesis, denitrification and uranium-reduction.
Experiments were performed with different anaerobic mixed cultures (including a sulfuroxidizing denitrifying mixed culture and a methanogenic sludge) at increasing initial
U(VI) concentrations. A high uranium-reducing activity could be observed in
250
methanogenic mixed cultures in the presence of H2 as electron donor, either in the
presence or absence of nitrate.
High activity towards uranium reduction was also
observed in experiments with acetate, but this was mainly attributed to the endogenous
substrates present in the methanogenic sludge. Uranium reduction was not observed
during denitrification with elemental sulfur or acetate. These results suggest that the
selection of electron donor effects the way nitrate impacts U(VI) reduction. H2 as electron
donor prevented nitrate form interfering with U(VI)-reduction.
Different levels of inhibition by U(VI) were observed depending on the microbial
process monitored. Acetoclastic methanogens were greatly inhibited with increasing
concentrations of U(VI). The 50% inhibiting concentration (IC50) was 0.16 mM U(VI).
The inhibition of denitrification was strongly dependent on the electron donor added.
U(VI) had an IC50 of 0.32 mM towards sulfoxidizing denitrifiers. A higher inhibition was
observed with H2- and acetate-utilizing denitrifiers (IC50 of 0.20 and 0.15 mM U(VI),
respectively). In contrast, the U(VI)-reducing activity in anaerobic sludge was not
inhibited, but instead stimulated up to concentrations of 1.0 mM. The inhibition studies
provide a better understanding of the potential microbial dynamics at uranium- and
nitrate- contaminated sites, so that conditions can be optimized in future bioremediation
systems.
Important implications for field applications of uranium bioremediation are
indicated by the research of this dissertation. Firstly, the findings reveal the potential of a
251
highly efficient uranium bioremediation in low-cost, minimum maintenance systems with
anaerobic granular sludge from methanogenic reactors. The granular sludge does not
require addition of external electron donor due to the continuous endogenous supply of
electron donor. Secondly, a remarkable enhancement of the U(VI)-reduction rate with Fe0
can be achieved with microorganisms in permeable reactive barriers based on Fe0 as the
reactive material. However, the specific mechanisms of U(VI) reduction with the
secondary products of biogenic Fe2+ need to be further studied. Finally, this work also
provides clues on the microbiological effects of U(VI) and nitrate in uranium
bioremediation systems under anaerobic conditions.
252
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