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Uranium isotope fractionation during adsorption, (co)precipitation and biotic reduction Duc Huy Dang, Breda Novotnik, Wei Wang, R. Bastian Georg, and Robert Douglas Evans Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b01459 • Publication Date (Web): 31 Oct 2016 Downloaded from http://pubs.acs.org on November 1, 2016

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Environmental Science & Technology

1

Uranium isotope fractionation during adsorption, (co)precipitation and biotic reduction

2 Duc Huy Dang1*, Breda Novotnik1, Wei Wang1, R. Bastian Georg2, R. Douglas Evans1,2

3 4 5

1

School of the Environment and 2 Water Quality Center, Trent University, 1600 West Bank

6

Drive, Peterborough, ON, Canada K9L 0G2

7 8

*Corresponding author. Tel: +1 705 748 1011 (ext. 7692).

9

E-mail: [email protected]

10

1

ACS Paragon Plus Environment


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11

ABSTRACT

12

Uranium contamination of surface environments is a problem associated with both U-ore

13

extraction/processing and situations in which groundwater comes into contact with geological

14

formations high in uranium. Apart from the environmental concerns about U contamination,

15

its accumulation and isotope composition have been used in marine sediments as a paleo-

16

proxy of the Earth’ oxygenation history. Understanding U isotope geochemistry is then

17

essential either to develop sustainable remediation procedures as well as for use in paleo-

18

tracer applications. We report on parameters controlling U immobilization and U isotope

19

fractionation by adsorption onto Mn/Fe oxides, precipitation with phosphate and biotic

20

reduction. The light U isotope (235U) is preferentially adsorbed on Mn/Fe oxides in an oxic

21

system. When adsorbed onto Mn/Fe oxides, dissolved organic carbon and carbonate are the

22

most efficient ligands limiting U binding resulting in slight differences in U isotope

23

composition

24

(δ238U=0.39±0.04‰). Uranium precipitation with phosphate does not induce isotope

25

fractionation. In contrast, during U biotic reduction, the heavy U isotope (238U) is

26

accumulated in reduced species (δ238U up to -1 ‰). The different trends of U isotope

27

fractionation in oxic and anoxic environments makes its isotope composition a useful tracer

28

for both environmental and paleo-geochemical applications.

(δ238U=0.22±0.06‰)

compared

to

the

DOC/DIC-free

configuration

29 30

1. INTRODUCTION

31

Uranium (U) is the heaviest naturally occurring element on Earth. Uranium

32

compounds (e.g., uranium trioxide, UO3) have been used for centuries in the production and

33

coloring of glass. More recently, applications for uranium have expanded to include nuclear

34

power generation and atomic weapons. The byproduct of nuclear fuel production (depleted

35

U) is currently used as ballast for ships and as counterweight for aircraft. In addition, other

36

anthropogenic activities (e.g. the extraction and processing of uranium) lead to uranium

37

mobilization and contamination of surface and ground waters, soils and sediments in many

38

parts of the world 1–3.

39 40

Previously, isotopic fractionation, i.e. differential behaviours of

235

U and 238U, was not

considered to be significant given the small differences in mass (ca. 1%) 4,5. However, recent 2


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41

advances in analytical techniques have revealed that there is considerable variation in uranium

42

isotope ratios in ores, reduced sediments, coral, manganese deposits, basalts, granites,

43

seawater, black shales, suboxic sediments, and ferromanganese crusts/nodules4–6. In addition,

44

U isotope ratios are affected by the process of enrichment during nuclear fuel production and

45

so areas which have been contaminated with either enriched uranium (accidental release from

46

nuclear power stations) or with depleted uranium (weapons testing sites and/or war zones)

47

will have distorted isotopic signatures (i.e. because of anthropogenic processes)

48

isotopic composition has subsequently been used, for example, in soil, surface and

49

groundwater samples to detect source inputs/discharge of natural and/or artificially altered

50

uranium to various environmental compartments 9–12.

7,8

. The U

51

Uranium isotope composition (238U/235U) also has been suggested as an indicator of the

52

oxygenation history of the Earth, the redox state of the ocean or to track U migration from

53

contaminated aquifers

54

thought to be one of the main factors controlling variations in δ238U making it a useful tracer

55

of oceanic redox conditions

56

with other natural materials such as volcanic rock, seawater and carbonate minerals, implies

57

that the redox transition from U(VI) to U(IV) at low temperature is the primary cause of

58

isotope fractionation 1,4,5,11,15,16. Biotic reduction processes lead to enrichment of the heavy U

59

isotope (238U) in reduced species (negative δ238Udissolved ~ -1 ‰ 5,17) while abiotic reduction of

60

U does not seem to induce significant U isotopic fractionation17. U isotope fractionation has

61

been observed between ferromanganese crusts and seawater, with an enrichment of the light

62

U isotope (235U) onto oxide surfaces (positive δ238Udissolved ~ 0.2 ‰

63

on a Mn oxide surface also has been found in a laboratory study using two U concentrations

64

(30 and 140 µM at pH 5 with N2 and CO2 sparging) 13. Under these conditions, similar values

65

of δ238Udissolved to those found on natural ferromanganese crusts were measured. However,

66

while U accumulation and U isotope fractionation are effectively promising tracers of

67

historical geochemical processes, the mechanism(s) of U fractionation, particularly in the

68

presence of complexing ligands are still not well documented.

1,13,14,15

. The biotic reduction of soluble U(VI) to insoluble U(IV) is

1,14

. Significant enrichment of 238U in reduced species compared

5,15

). Uranium adsorption

69

The continuous increase in the global demand for energy together with pressures to

70

provide sustainable energy sources, necessitate the development of eco-friendly uranium

71

extraction techniques and uranium remediation processes. The latter frequently involves U 3



72

immobilization via processes such as (bio)adsorption, (bio)precipitation and (bio)reduction

73

1,12,18

74

efficiency of remediation technology relies on the biogeochemical and transport processes

75

governing uranium mobility as well as the stability of the immobilized U species 1,19. In fact,

76

bioreduction can be an efficient removal mechanism leading to a decline in U concentrations

77

in ground water by an order of magnitude

78

depends strongly on other environmental factors (e.g. redox variation, chemical species).

79

Manganese and Fe exert a dual role on the two valence states of U. Mn/Fe oxides are efficient

80

scavengers of U(VI)

81

and Mn oxides can act as an oxidant inducing U(IV) (UO2(s)) oxidative dissolution and this

82

oxidation is fostered by carbonate and organic chelators (natural organic matter or bacterially

83

secreted substances)

84

uptake on oxide surfaces at U concentrations below 10 µM but at higher U concentrations (up

85

to 130 µM, where U-P should precipitate) the P-Fe binding results in higher dissolved U

86

concentrations compared to Fe-free conditions

87

interactions in natural environments strongly affect the efficacy of U remediation processes.

88

In these systems, monitoring only U concentrations does not appear adequate to assess the

89

main mechanisms controlling U mobility. However, variations in U isotope ratios, i.e. isotopic

90

fractionation, could provide additional information on the processes that ultimately control U

91

solubility and mobility.

. All involve uranium transformation to a less mobile form. However, the long-term

20,21

1,17

. However, the stability of reduced U minerals

in surface environments. However, in subsurface environments, Fe

22–24

. In addition, the interaction among U, P and Fe oxides enhances U

20,25

. In summary, complex biogeochemical

92

In this study, we have conducted batch experiments to determine U immobilization and

93

U isotope fractionation during adsorption onto Mn/Fe oxides, into phosphate precipitates and

94

during U bioreduction. We monitored also the influence of ligands (dissolved

95

inorganic/organic carbon (DIC/DOC)) and co-precipitating elements (Ca, P) that are

96

commonly present in aquatic systems on U immobilization by adsorption, precipitation, and

97

bioreduction in the presence of an electron donor.

98 99

2. MATERIALS AND METHODS

100 101

2.1. Adsorption experiments

4


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102

Batch adsorption experiments were conducted using solutions containing varying

103

concentrations of U (1 to 100 µM), plus different concentrations of DIC (primarily as the

104

bicarbonate species at a pH used for these experiments), DOC and/or Ca in contact with δ-

105

MnO2 or goethite. The solid substrates were synthesized using established methods

106

Briefly, to produce geothite, a solution of 5 M KOH was gradually added to a 1 M Fe(NO3)3

107

solution to precipitate ferrihydrite, which was then diluted in ultrapure (18MΩ) water and

108

kept at 70ºC for 60 h for crystallization to goethite. The precipitate obtained was centrifuged

109

(3500 g for 10 min) and dialyzed (Spectralab 1000 Da) four times to remove excess ions.

110

Following centrifugation, the goethite was collected and then freeze-dried. The specific

111

surface area of this synthesized goethite is assumed to be approximately 39.9 m2 g-1 25. δ-

112

MnO2 was prepared according to the redox method

113

KOH was added gradually to a Mn(NO3)2 solution. The Mn oxide obtained was centrifuged

114

and rinsed with 1M NaCl solution before being dialyzed. The δ-MnO2 was then collected by

115

centrifugation and freeze-dried.

25,26,21,27

.

21,27

. Briefly, a mixture of KMnO4 and

116

Duplicate batch experiments were conducted in 250 and 50 mL (δ-MnO2 and goethite,

117

respectively) HDPE bottles, which were previously cleaned with 10% HNO3 (Trace Metal

118

Grade) and then rinsed twice with ultrapure water.

119

concentrations: 0, 1, 5, 10, 50, 100 µM) was mixed with either δ-MnO2 or goethite at a

120

solid/liquid ratio of 0.1 and 0.5 g L-1, respectively. The U solution was prepared from an

121

isotopic reference material, IRMM 184 (European Commission-Directorate General Joint

122

Research Center), containing a specific

123

experimental solutions was kept constant at ca. 7.1 with MOPS buffer (pKa = 7.2, final

124

concentration of 15 mM) and the ionic strength maintained at 10 mM with NaNO3. Calcium,

125

carbonate and organic matter are the main chemical species forming complexes with U in

126

aquatic environments 29. In order to assess the influence of DIC (NaHCO3, Acros Organics),

127

DOC (Suwannee River Fulvic Acid Standard II, International Humic Substances Society) and

128

Ca (CaCl2, Acros Organics) on U fractionation, different concentrations of DIC (0, 1 and 5

129

mM), DOC (0, 10 and 20 mg C L-1) and Ca (0, 1 and 5 mM in the presence of 1 mM DIC)

130

were used for the batch experiments.

238

A solution of U (at different

U/235U ratio of 137.679

28

. The pH of the

131

In a previous study U adsorption on Mn oxides was shown to be rapid with isotope

132

fractionation being stable after 2h to 48h13. However U adsorption onto hematite reached only 5



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30

133

~67% of total adsorption after 24h and gradually increased up to 5 days

. Thus, our batch

134

experiment ran for 5 days, after which the water was collected by filtration (0.2 µm, Nylon,

135

Sartorius) and then acidified (double-distilled Trace-Metal-Grade HNO3) prior to measuring

136

total U and U isotope ratios. Prior to DIC/DOC analyses, NaN3 (final concentration of 1 mM)

137

was added to the filtered water to prevent bacterial activity.

138 139

2.2. Phosphate precipitation experiments

140

Uranium phosphate precipitation contributes to U immobilization

31

. However,

141

phosphate availability is limited by other elements (e.g. Ca by hydroxyapatite precipitation 32

142

and/or Fe by surface scavenging and precipitation

143

with phosphate were conducted by mixing a U solution (1 or 100 µM, IRMM 184) with

144

different concentrations of phosphate (1 µM to 50 mM; K2HPO4, which was assumed to be

145

soluble reactive phosphate (SRP)) to a final volume of 15 mL. The experimental pH was

146

stabilized by MOPS buffer (pKa = 7.2, final concentration of 15 mM) and the ionic strength

147

was maintained at 0.1 M with NaNO3. To assess the influence of DIC and Ca, different

148

configurations were set up: (i) DIC only (0 and 2 mM), (ii) Ca only (0 and 10 mM) and (iii)

149

combination of DIC and Ca. The solutions were shaken for 24 hours, centrifuged (15 min,

150

4000g), resulting in a yellow pellet of the U-P precipitate. In the experiments with Ca, a milky

151

white Ca-P precipitate was formed. The supernatant water was filtered (0.2 µm, Nylon,

152

Sartorius) and acidified (double-distilled Trace-Metal-Grade HNO3) prior to measuring total

153

U concentrations and U isotope ratios. The pellet was dissolved in 0.5% HNO3 to determine U

154

recovery.

25

). Experiments to assess U precipitation

155 156

2.3. Biotic reduction experiments

157

Shewanella algae BrY (Culture Collection, University of Göteborg, Sweden) were

158

grown aerobically in Luria Bertani (LB) medium at room temperature to late stationary phase

159

(16 h) then collected by centrifugation and washed with a medium containing 20 mM PIPES

160

buffer and 30 mM sodium bicarbonate (pH 6.8). For U(VI) biotic reduction, assay cells were

161

suspended to an optical density (OD600) of 0.5 ± 0.05 in 150 ml of the previous simple

162

medium (buffer and bicarbonate) containing ca. 6 μM U(VI) and lactate at various

163

concentrations (0, 5, 20, 30 and 60 mM). All experiments for U(VI) bioreduction were 6


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164

conducted at 20 ºC under anoxic conditions (i.e., purged with N2 for 30 min) using anaerobic

165

culture bottles (250 mL, Schott) with screw caps having a silicone o-ring and blue butyl

166

rubber stopper. All solutions were filter-sterilized. Samples were withdrawn every day for 4

167

days through the butyl rubber stopper using a needle and syringe.

168 169

2.4. Determination of total U and U isotope composition

170

The concentration of U was determined by QQQ-ICP/MS (Agilent 8800) using In as an

171

internal standard; accuracy was checked using a ground water reference material (ES-H-1,

172

EnviroMat). The DOC/DIC analyses were performed on a TOC-VCSH analyzer (Shimadzu);

173

accuracy was checked using a river water reference material (PERADE-09, Environment

174

Canada).

175

For U isotope analyses, chemical separation of U from the matrix was performed using

176

TRU resin (100-150 µm, Eichrom). A chromatography column was loaded with 0.5 mL of the

177

resin which was sequentially rinsed with 10 mL of 0.1 M HCl/0.3 M HF then 10 mL of 0.2 M

178

HCl. The TRU resin was conditioned with 1.5 M HNO3 before approximately 600 ng of U

179

from the sample were loaded on the column. A double spike technique with a pre-mixed

180

233

181

fractionation and potential isotope fractionation on the column 4,5. Approximately 12 ng each

182

of

183

resin was rinsed with 7 mL (i.e. 4 x 0.5 + 5 mL) of 1.5 M HNO3, 5 mL of 3M HCl then 10

184

mL of 1 M HCl to remove most matrix elements from the column. The final elution of U was

185

performed using 6 mL (i.e. 2+4 mL) of 0.1 M HCl/0.3 M HF solution. The recovery was in

186

the range of 98-100 %.

U/236U solution (IRMM 3636b) was used for internal correction of instrumental mass 233

U and

236

U (IRMM 3636b) were loaded on the TRU resin along with the sample. The

187

The U isotope measurements were made using MC-ICP/MS (Nu Plasma II) at the Water

188

Quality Center (Trent University, Canada). The Nu Plasma II was equipped with 16 fixed

189

Faraday detectors, a Ni sample cone (ES Dry Plasma FG9, Nu instruments) and a Ni skimmer

190

cone (ES Dry Plasma HS1-7, Nu instruments). The measurement of 233U, 235U,

191

was performed on the L3, L1, Ax and H2 cups respectively. Samples were introduced using a

192

Cetac Aridus II combined with a PFA nebulizer. The U isotopes were measured for 80-100

193

cycles following the background measurement and peak centering at the block start. The

194

instrument was tuned with a solution of 60 ppb U (IRMM 184) giving a ~30 V signal on 238U. 7


236

U and 238U


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195

The signals for 233U, 235U and 236U were ca. 600, 200 and 600 mV, respectively. The 235U and

196

238

197

spike correction was performed using the exponential law. Each sample was bracketed by two

198

double-spiked standard solutions (IRMM 184) and the U concentrations in standards were

199

adjusted to that of the samples ±10%. Washout between samples was achieved using 0.1 M

200

HCl/0.3 M HF solution. Uranium isotope variations in the samples are reported relative to that

201

of the standard (IRMM 184), using the equation:

U contributions from the double spike solution (IRMM 3636b) were corrected. The double

δ

202

The

238

U/ U

U =

− 1 × 1000 U/ U

U/235U ratio in the IRMM 184 was certified to be 137.697±0.042

28

. An inter-

203

calibration campaign among several laboratories worldwide resulted in a ratio of

204

137.683±0.020 33,34. Our measured238U/235U ratio was 137.677±0.040 (n=73) , which is within

205

the values previously reported 28, 21, 22.

206 207 208

3. RESULTS AND DISCUSSION

209

reduction

210

3.1.1. Adsorption experiments

3.1. Uranium sequestration by adsorption, precipitation with P and biotic

211

Uranium adsorption onto Mn and Fe oxides shows a similar behavior (Fig. 1A). Both

212

Mn and Fe oxides bind close to 100% of U in solution at the lower U concentrations (up to 10

213

μM) and up to 70% of the 100-μM U solution. At the highest U concentration (100 μM), U

214

adsorption (Fig. 1A) reaches the theoretical binding site saturation for both Mn and Fe oxides

215

(0.56 and 0.12 meq g-1 for Mn and Fe, respectively 21,25).

216

The results for U adsorption onto Mn oxide and Fe oxide in the presence of Ca, DIC and

217

DOC (Fig 2 and S1, respectively) indicate that a DIC concentration of 1 mM increases the U

218

adsorption onto Mn oxides only slightly (Fig. 2A), while for Fe oxide, binding is independent

219

of the initial U concentration (93.2 ± 1.6 %, n = 5, Fig. S1A). At a DIC concentration of 5

220

mM, U adsorption onto both Mn and Fe oxides is reduced to ca. 40 %. These observations are

221

in agreement with Waite et al.

222

(e.g. low carbonate concentration) is favored over the coordination of uranyl with two or more

35

; the 1:1 uranyl-carbonate complex binding on ferrihydrite

8


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223

carbonate ligands onto oxides. The presence of calcium inhibits U adsorption (Figs. 2B, S1B)

224

36

225

oxide surface. Calcium adsorbs weakly onto Fe oxide 37 but strongly onto birnessite at neutral

226

pH 38. Calcium forms not only a stable dissolved calcium uranylcarbonate complex, but also it

227

adsorbs onto Mn oxides, both processes leading to decreased U binding.

and the effect is more noticeable for Mn oxide, probably due to Ca adsorption onto the

228

In the presence of DOC, U behavior is similar for Fe and Mn oxides (Figs. 2C, S1C); at

229

lower U concentrations, an increase in DOC concentration corresponds to a significant

230

decrease in U adsorption. However, the impact of DOC becomes less noticeable when the U

231

concentration exceeds the binding capacity of the DOC. For comparison, Trenfield et al. have

232

calculated that 88% of U is bound by SRFA at 10 mg C L-1 and 95 μg L-1 of U

233

conditions, a plateau for Uads of ~70 % onto Mn oxides was observed (Fig. 2C). No change in

234

DOC concentration was observed over the course of the experiment (data not shown).

39

. In our

235 236

3.1.2. U-P precipitation experiments

237

The precipitation of U increases sharply when P reaches a critical concentration (Fig

238

1B); the SRP concentration leading to removal of 50% of U is ca. 1 and 10 μMSRP for initial

239

U concentrations of 1 and 100 μMU, respectively. The presence of DIC and Ca maintains the

240

U in solution (Fig. 3A, B) compared to the DIC- and Ca-free treatments, except for the 100-

241

μMU experiment where Ca enhances U removal (when comparing the 10-mM Ca

242

configuration to the DIC- and Ca-free system, Fig 3B). This difference could be caused by the

243

(co)precipitation of Ca-P-U when U is present at a high concentration. When both DIC and

244

Ca are present, the effect seems to be additive, preventing U-P precipitation by stabilizing U

245

in a calcium uranylcarbonate complex while reducing available P (Fig. 3A). However, at a

246

high U concentration (100µM), coprecipitation of U with Ca-P precipitate could be possible

247

(Fig. 3B).

248

The U-P-Ca-DIC system was modeled with PHREEQC (version 2.18.3, 40) similar to an

249

approach previously described 41. Briefly, the chemical speciation and the mineral saturation

250

index in mixtures consisting of various concentrations of U (1 to 100 μM) and P (0.1 μM to

251

100 mM), with or without Ca (10 mM) and DIC (2 mM) was first calculated. Second, using

252

the EQUILIBRIUM_PHASE module, the solid/liquid distribution of U was simulated by

253

allowing mineral precipitation when the saturation index reached a positive value. From the 9



254

first calculations, the main minerals that can potentially precipitate are chernikovite

255

(UO2HPO4.4H2O), uranyl orthophosphate ((UO2)3(PO4)2) and autunite (Ca(UO2)2(PO4)2). The

256

precipitation of these minerals has been previously demonstrated experimentally and by

257

modeling 25.

258

The model produced from PHREEQC was compared to our experimental data, under

259

four conditions (Ca- and DIC-free; Ca only; DIC only; Ca and DIC; Fig. S2). The results

260

indicate that in the Ca- and DIC-free systems, the main dissolved U species are uranyl

261

hydroxide and uranyl phosphate and they are dominant when SRP is not at the critical

262

concentration (Fig. 4A). The critical concentration of SRP leading to efficient U trapping

263

ranges from (10-6 to 10-4 M) as the U concentration increases from 1 to 100 µM. Over this

264

critical SRP concentration, uranyl phosphate is the main U-P mineral that is

265

thermodynamically favored. In the presence of DIC, the formation of uranyl carbonate

266

complexes (UO2(CO3)x2(1-x), prevent U-P precipitation until the SRP concentration reaches a

267

higher critical concentration ([SRP] = 10-5 to 10-3 M) than in the Ca- and DIC-free system

268

(Figs. 4A and C). The main uranium precipitate is autunite (Ca(UO2)2(PO4)2, Fig. 4D).

269

Uranium precipitation with SRP does not induce any change in U isotope composition,

270

as the U-P interaction does not favor any light or heavy U isotopes. As for U adsorption onto

271

Mn/Fe oxides, U speciation and mobility are tightly linked. Complexation with either organic

272

or inorganic ligands is efficient in preventing U precipitation whereas the presence of Ca

273

either keeps U in solution (via calcium uranylcarbonate complexes) or enhances U trapping in

274

a Ca-phosphate precipitate.

275

3.1.3. Biotic reduction

276

The results of the control experiment (0 mM lactate, Fig. 1C) indicate that 80% of U

277

remains in solution over the course of the experiments. This U removal could be related to U

278

retention on the outer membrane of the bacteria or, because Shewanella algae is a gram-

279

negative bacterium, there may be U penetration through the outer membrane into the

280

periplasm, where various C-type cytochromes involved in electron transfer are present 18,42,43.

281

By increasing the lactate concentration from 5 to 30 mM, U removal is enhanced (5% to 70%,

282

respectively of U was further removed relative to the control) which is related to U reduction

283

by supplying an electron donor to the system. Moreover, a high lactate concentration (60 mM,

284

Fig. 1C) decreases the U reduction, probably due to the mechanism of substrate inhibition 44. 10


Page 10 of 29

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285

The high concentration of substrate could inhibit bacterial growth, distorting its metabolism.

286

The addition of the electron donor (acetate, lactate) has been proposed as an alternative to

287

traditional U remediation (adsorption or precipitation) by boosting the bacterial reductive

288

activity 1,18.

289 290

3.2. Uranium isotopic composition

291

Uranium adsorption onto Mn oxide shows a significant and stable fractionation

292

independent of total U (Fig. 6A). The occurrence of other ions and compounds seems to affect

293

U fractionation, where DIC-, DOC- and Ca-free configurations (i.e. U is not bound to these

294

ligands) show a systematically higher δ238U (δ238U = 0.39±0.04‰) than the configurations

295

with DIC, Ca and DOC (δ238U = 0.22±0.06‰). The latter value is close to the U fractionation

296

during adsorption onto birnessite in a carbonated system (0.22 ±0.09‰

297

ferromanganese crust relative to seawater (0.24‰

298

induces fractionation that is similar to that observed with Mn oxide (without addition of

299

ligands) at high total U concentrations (50 and 100 µM) where more than 50% of U is bound

300

(δ238U = 0.34±0.09 ‰). At lower U concentrations, U fractionation decreases toward the

301

initial U isotope composition (Fig. 6B). δ238U appears to be systematically higher in the DIC-,

302

DOC- and Ca-free systems compared to those with Ca, DIC and DOC (Fig. 6A and B).

303

Uranium precipitation with SRP does not induce any fractionation relative to the initial U

304

isotope composition (Fig. 6C). However, the biotic reduction of U induced a depletion of 238U

305

in solution (negative δ238Udissolved). In the first 24h of the experiment, when lactate is supplied,

306

removal of dissolved U (up to 20 %, Fig. 1C) is approximately the same as in the control

307

(without lactate). This is presumably due to biotic adsorption (section 3.1.3.) and is not

308

accompanied by U fractionation (Fig. 4D). Uranium isotope fractionation is clearly observed

309

when the majority of U is precipitated after the first 24h by bioreduction (Figs. 1C and 6D).

5,15

13

) and onto a

). Uranium adsorption onto Fe oxide

310 311

3.2.1. Uranium fractionation during adsorption

312

Uranium fractionation during adsorption leads to an accumulation of the light isotope

313

onto Mn/Fe oxides (positive δ238Udissolved, Figs. 6A and B). This observation is consistent

314

with U fractionation recorded in marine ferromanganese nodules and crusts

315

experiments 13. A very similar behavior has been observed also for Mo 11


45–47

5,15

and by lab

. The mechanism


Page 12 of 29

316

of element fractionation during adsorption could be either (i) equilibrium isotope exchange

317

between aqueous and adsorbed species or (ii) isotope exchange between two or several

318

aqueous species followed by one of these aqueous species binding specifically to the oxide

319

surface sites. Barling et al. 46 suggest that the breakage of the >S-OH (S being the adsorption

320

site on the oxide) forms a >S-O-Mo bond and the Mo-O bond in >S-O-Mo (adsorbed Mo) is

321

weaker than the dissolved Mo-O. Furthermore, the dissolved

322

than the

323

preference of light Mo isotope onto the adsorption surface and enrichment of the heavy Mo

324

isotope in solution.

98

Mo-O due to higher vibrational energy

47

95

Mo-O bond is more fragile

. Altogether, this mechanism leads to a

325

For U, a similar mechanism involving the equatorial U-O single bond could be

326

plausible. However, because U is a much larger atom than Mo, the nuclear field shift means

327

that the electronic energy of the light isotope lies lower than that of the heavy isotope48. Like

328

Mo, the difference in energy between >S-O-U and dissolved U-O should result in isotope

329

fractionation during adsorption. However, the smaller differences in mass and energy

330

discrepancy between the light and heavy isotopes within the dissolved U-O bond should

331

attenuate the isotope fractionation factor (δ238U ~ 0.2 ‰ compared to that of δ98Mo ~ 2 ‰).

332

The presence of different ligands in the media (Ca, DIC, DOC) seems to affect U

333

isotope fractionation onto Mn/Fe oxides (Fig. 6A and B) as well as the U binding capacity to

334

the oxides (Figs. 2 and S1). Adsorption of U on oxides induces significant U fractionation

335

(0.39±0.04‰) while lower U fractionation is observed in the presence of ligands. In addition,

336

U fractionation as a function of percentage of dissolved U (Fig. S3) corresponds to the

337

fractionation model in a closed system equilibrium, as was observed for Mo fractionation

338

during adsorption on Mn oxide

339

and independent of the initial U concentrations (Fig. 6A), unlike the experiments with Fe

340

oxide (Fig. 6B). The difference could be caused by the adsorption mechanism of U onto Mn

341

and Fe oxides; U adsorption onto Mn oxide at pH below 8 has been demonstrated by EXAFS

342

to be mainly bidentate-mononuclear (>(MnO)2-UO2, model II, Fig 5B)

343

adsorption onto Mn oxide 46, U binding on the Mn oxide surface leads to a constant U isotope

344

fractionation, indicating a single mechanism. The geometry calculation from binding structure

345

and bond length given by Wang et al.

346

U atoms bound on Mn oxide is ca. 5.18 Å. Given that the U and equatorial O (U-Oeq) distance

46

. However, uranium fractionation onto Mn oxide is stable

21

21

. Similar to Mo

(see Table 1) suggests that the distance between two

12


Page 13 of 29

347 348 349


is ca. 2.46 Å

21,30,49

, two equatorial shells of each U overlap. Therefore, we suggest that the

235

lighter U isotope ( U) with a lower electronic energy than partitioning of

235

238

U

48

could favor preferential

U to the adsorption sites.

350

On the other hand, it has been demonstrated that U adsorbs onto hematite by bidentate-

351

mononuclear (model I, Fig. 5B) and bidentate-binuclear (model III, Fig. 5B) binding30. At pH

352

below 8, the bidentate-binuclear mode is favored when the initial U concentration is above 10

353

μM, as demonstrated by the U-U bond (EXAFS fitting 30). In our experiments with Fe oxide,

354

δ238U increases with initial concentration of U (Fig. 6B). This observation suggests that U

355

adsorbs onto Fe oxide by two mechanisms (Models I and III, Fig. 5B) leading to U isotope

356

fractionation in one configuration and less in the other (Fig. 6B). A similar geometry

357

calculation was performed using EXAFS data for U adsorbed onto hematite and ferrihydrite

358

30,35

359

concentration), the distance between two U atoms is ca. 7.42 Å and there is no overlapping of

360

the equatorial shell (r ~ U-Oeq = 2.46 Å); uranium fractionation is thus not significant (Fig.

361

6B). Alternately, in the bidentate-binuclear configuration (model III, Fig. 5B, high initial U

362

concentration), the two U atoms share the equatorial shell (an intermediate O atom) and the

363

distance between themis ca. 3.9 Å 30. The U isotope fractionation is then in the same range as

364

seen for Mn oxide.

(see Table 1). In the bidentate-mononuclear configuration (model I, Fig. 5B, low initial U

365

Thus, it would appear that U isotope fractionation is highly dependent on the adsorbate

366

structure, mostly the density of the surface binding sites, and on the concentration of U in the

367

initial solution. Our results have demonstrated that U isotope fractionation during adsorption

368

onto Mn oxides in the presence of DIC, Ca and DOC is less significant (δ238U = 0.22±0.06‰)

369

than during U adsorption (δ238U = 0.39±0.04‰) in the absence of these ligands (Fig. 6A),

370

which matches the value reported in marine ferromanganese crusts (0.22-0.24‰)

371

could be a coincidence due to the equilibrium fractionation mechanism (Fig. S3). However, if

372

further studies confirm a constant difference between the two configurations previously

373

detailed, this difference in δ238U (∆δ238U ~ 0.2‰) during U adsorption could be used to

374

indicate the U speciation. Further experiments coupling structural studies (e.g. EXAFS) and U

375

isotope composition should provide additional insight.

376 377

3.2.2. U fractionation by biotic reduction, effect of electron donor supply 13


5,13,15

. That


378

Page 14 of 29

U(VI) reduction to U(IV) by biotic or abiotic mechanisms has been demonstrated to be 17,18,50

379

a main process leading to U immobilization in anoxic environments

. The biotic

380

reduction mechanism is based on a microbial one-electron exchange, to reduce U(VI) to

381

U(V), and then a disproportionation of U(V) to U(IV) and also to U(VI) 51. The U reduction is

382

catalyzed by enzymatic activities (multiheme C-type cytochromes) and the reductase should

383

be found either on the outer cell membrane or in the periplasm of the bacteria42.

384

In our study, when the lactate concentration was 30 mM or less, U removal (ca. 20% of

385

total U) was systematically recorded after 24 h; this was not significantly different than the

386

control (0 mM lactate) (Fig. 1C). This first-step U removal does not induce U fractionation

387

and could be due to adsorption onto the cell membrane or U uptake into the periplasm to

388

access U reductase. The fact that at low lactate concentrations, we observe results similar to

389

the control, where an electron donor is not supplied, would suggest that bioadsorption is the

390

only mechanism occurring during the first 24h of the experiment but not bioreduction. A

391

similar process was observed for Np and G. sulfurreducens

392

and when lactate is present in the solution, reductase is activated, and U removal increases; it

393

is highest at a lactate concentration of ca. 20-30 mM and is accompanied by U isotope

394

fractionation (Figs. 1C and 6D). The heavy U isotope (238U) is removed from the dissolved

395

fraction leading to a negative δ238Udissolved.

396

51

. However, after the first 24h

For heavy elements such as Tl, Hg, U, isotope fractionation during reduction results 48,52

397

mainly from the nuclear field shift

, caused by the difference in size and shape of the

398

isotope nuclei. The contribution of conventional mass-dependent isotope fractionation would

399

be less significant for these elements because of their small relative mass differences

400

Previous calculations for Hg and Tl have demonstrated that the mass-dependent mechanism

401

would induce a fractionation of 0.5-1 ‰ compared to 3 ‰ for the nuclear volume effect 53.

402

The nuclear field shift (NFS) theory predicts that the heavy isotope would be enriched in the

403

chemical species with the smallest electron density at the nucleus 205

52,53

.

48,53

. For example, when

404

thallium (with two isotopes) is fractionated, heavy Tl (

405

Tl(III) (Tl3+: [Xe] 4f14 5d10) whereas light Tl (203Tl) is accumulated in reduced species, Tl(I)

406

(Tl+: [Xe] 4f14 5d10 6s2) 52,54. Similar to Tl, Hg reduction (Hg2+/Hg0) results in enrichment of

407

heavy

408

(e.g. an increase in electron density at the nucleus). In other words, oxidized species of Hg

202

Tl) is enriched in oxidized species,

Hg in Hg2+ 53. The Hg and Tl reduction corresponds to a gain of two 6s electrons

14


Page 15 of 29


409

and Tl which hold smaller electron densities at the nucleus are enriched in heavy isotopes. In

410

contrast, the heavy U isotope (238U) is accumulated in reduced U(IV) (Fig. 6D) 2+

2+

5,17,55

. This

3

1

411

discrepancy comes from the molecular orbit of U(VI)O2 , built from U ([Rn] 5f 6d ) and 2

412

axial O48 compared to that of U(IV) (U4+: [Rn] 5f1). Unlike the reduction of Hg and Th,

413

U(VI)/U(IV) reduction results from the loss of 2 5f electrons, accompanied by a decrease in

414

electronic density at the nucleus

415

observed U isotope fractionation which is opposite to Tl and Hg.

52

. This difference explains the NFS effect inducing the

416 417

4. ENVIRONMENTAL IMPLICATIONS

418

This study shows that Mn/Fe oxides, phosphate and bacterial activity can affect the fate

419

of U in environmental compartments. In situ remediation strategies involving these

420

parameters to control U mobilization from contaminated sites have been previously proposed

421

and considered

422

implications. For example, MnII and FeII oxidation when discharged from ground water to

423

surface water can generate an oxide reactive barrier immobilizing contaminants

1,13,18,56

but further studies are required to better evaluate long-term 21

but Mn

19,57

424

oxides can also enable reoxidization or dissolution of solid U(IV) to soluble U(VI)

. The

425

presence of U-binding ligands can inhibit U immobilization by adsorption and precipitation

426

but also influence the final products of bioreduction

427

chemistry of the remediation target site is essential for adequate planning.

56

. Understanding the environmental

428

Uranium isotope composition could be applied to track geochemical processes or to

429

better evaluate U remediation procedures1. However, no measurable changes in U isotopic

430

composition were reported as a result of desorption/adsorption of U from particles in

431

groundwater 58 and U precipitation with phosphate (section 3.2.). Thus care must be exercised

432

in the interpretation of U isotopic fractionation. Also, the preferential precipitation of the 238U

433

isotope during biostimulation is confirmed to be a specific fingerprint of biotic reduction.

434

However, it is important to recognize that in aquatic environments U removal indicated by a

435

change in isotopic composition from biotic reduction may not be the only mechanism of

436

removal, as the adsorption/desorption in groundwater and precipitation with phosphate do not

437

induce U isotopic fractionation. In addition, the small but significant difference in U isotopic

438

composition during U adsorption onto Mn/Fe oxide in the presence and absence of DIC/DOC

439

(δ238U=0.22±0.06‰ and 0.39±0.04‰, respectively) could be useful to trace U speciation. To 15



440

summarize, the opposite fractionation of U isotopes observed, during adsorption onto oxides

441

(positive δ238U) versus biotic reduction (negative δ238U), makes U isotope composition useful

442

as a tracer for constructing models to predict redox variation.

443 444

ACKNOWLEDGEMENTS:

445

This work was supported by a Canadian NSERC (Natural Sciences and Engineering Research

446

Council) Collaborative Research and Development Grant to RDE which also funded DDH’s

447

and BN’s postdoctoral fellowships. The authors wish to thank Dr. Hayla Evans for English

448

correction and manuscript revision.

449 450

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613

20


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Abstract Graphic 61x47mm (150 x 150 DPI)



Figure 1: (A) Amount of adsorbed U and binding site occupancy (%) onto Mn (circles) and Fe (squares) oxides as a function of total initial U. (B) Dissolved U remaining after precipitation with phosphate as a function of P/U molar ratio. (C) Temporal variation of dissolved U in the biotic reduction experiments with different concentrations of lactate as an electron donor supply. Fig. 1 75x25mm (300 x 300 DPI)


Page 22 of 29

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Figure 2: Percentage of U adsorbed onto Mn oxides at various concentrations of DIC (A), Ca at 1mM DIC (B) and DOC (C). The error bars represent standard deviation of the analysis and duplicate experiments. Figure 2. 87x33mm (300 x 300 DPI)



Figure 3: Influence of DIC and Ca on U-P precipitation. The graphs represent dissolved U remaining after precipitation with phosphate within a P concentration gradient. Initial U concentrations are 1 µM (A) and 100 µM (B). Figure 3. 119x67mm (300 x 300 DPI)


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Figure 4: Simulation of dissolved U remaining (A and C) and principal precipitate minerals (B and D) after precipitation with P as a function of initial SRP and U concentrations. Two configurations are Ca- and DICfree (A and B) and 10 mM Ca-2 mM DIC (C and D). The main dissolved species as well as minerals are shown upper on the graphs. 213x197mm (300 x 300 DPI)



Figure 5: (A) Ball-and-stick representation of the dissolved UO22+ (left) and dissolved uranyl carbonate complex (right) with the two axial O perpendicular to the equatorial plane which are not shown. (B) Postulated models of U adsorption on Fe and Mn oxides. See text for more details. * value is calculated from structure geometry ** value is assumed as a double of Mn-Mn or Fe-Fe distance. Figure 5A. 265x159mm (96 x 96 DPI)


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Figure 5: (A) Ball-and-stick representation of the dissolved UO22+ (left) and dissolved uranyl carbonate complex (right) with the two axial O perpendicular to the equatorial plane which are not shown. (B) Postulated models of U adsorption on Fe and Mn oxides. See text for more details. * value is calculated from structure geometry ** value is assumed as a double of Mn-Mn or Fe-Fe distance. Figure 5B. 328x437mm (96 x 96 DPI)



Figure 6: Variation of U isotopic composition (δ238U relative to IRMM 184) as a function of total initial U concentration or percentage of dissolved U during adsorption on Mn (A) and Fe (B) oxides, precipitation with P (C) and biotic reduction (D). The dashed line in (A and B) represents the value reported in the literature for U isotope fractionation during adsorption onto marine ferromanganese crusts and nodules 5,15. See text for more details. The symbol size in D is proportional to sampling time (0 to 96 h). The error bars represent standard deviation of the analysis and duplicate experiments. 62x17mm (300 x 300 DPI)


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Table 1: Summary of U adsorption experiments on Mn/Fe oxides as well as EXAFS Spectra Fitting Parameters

Ferrihydrite Hematite

Goethite Bio-MnO2 δ-MnO2 Birnessite

U-Oax

U-Oeq

Distance (Å) U-C U-Me

1.79

2.35

3.42

6-7

1000 100 10-35 100 353500 350

4.3, 8.2

22

1.79

22

1.79

pH

Me/U

7.0 4.5-6.5 6.4-6.7 8-8.5 4

4.2, 8.4, 9 5

1.8

2.31 and 2.49

1.8

1.77

2.3-2.41 2.29 and 2.46 2.3 and 2.47 2.39

2.9

2.9

3.42-3.46

3.45 and 4.3

U-U 3.9 3.9

Ref

Me-Me

29

Monomeric complex Multimeric complex Multimeric complex

26

-

Monomeric complex

17

-

Monomeric complex

17

2.9, ,2.97, 3.36, 3.71

26 26

2.92

3.45 and 4.32

-

Monomeric complex

18

2.92

3.41 and 4.25

-

Monomeric complex

18

3.39

-

Monomeric complex

8

* calculated distance from structure geometry


2.59*

Uranium Isotope Fractionation during Adsorption ... - ACS Publications

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