Quantifying Technetium and Strontium Bioremediation Potential in

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Quantifying technetium and strontium bioremediation potential in flowing sediment columns. Clare L Thorpe, Gareth T.W. Law, Jonathan R. Lloyd, Heather A Williams, Nick Atherton, and Katherine Morris Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b02652 • Publication Date (Web): 02 Oct 2017 Downloaded from http://pubs.acs.org on October 5, 2017

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Environmental Science & Technology

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Quantifying technetium and strontium

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bioremediation potential in flowing sediment

3

columns.

4

Clare L. Thorpe1†, Gareth T. W. Law1/2, Jonathan R. Lloyd1, Heather A. Williams3, Nick

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Atherton4, Katherine Morris1*.

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1

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Environmental Science, School of Earth and Environmental Sciences, The University of

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Manchester, Manchester, M13 9PL, UK.

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2

Research Centre for Radwaste Disposal and Williamson Research Centre for Molecular

Centre for Radiochemistry Research, School of Chemistry, The University of Manchester,

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Manchester, M13 9PL, UK.

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3

Nuclear Medicine Centre, Manchester Royal Infirmary, Manchester, M13 9WL, UK.

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4

Sellafield Ltd., Land Quality, Sellafield, Seascale, Cumbria, CA20 1PG, UK.

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* Corresponding author: [email protected]

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†

15

Engineering, University of Sheffield, Sheffield, S1 3JD, UK.

Current Address: Immobilization Science Laboratory, Department of Materials Science and

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Radioactively contaminated land, metastable technetium-99, strontium, bioreduction, gamma

18

camera imaging.

ACS Paragon Plus Environment

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Abstract

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The high-yield fission products 99Tc and 90Sr are found as problematic radioactive contaminants

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in groundwater at nuclear sites. Treatment options for radioactively contaminated land include

22

bioreduction approaches and this paper explores

23

range of biogeochemical conditions stimulated by electron donor addition methods. Dynamic

24

column experiments with sediment from the Sellafield nuclear facility, completed at site relevant

25

flow conditions, demonstrated that Fe(III)-reducing conditions had developed by 60 days.

26

Sediment reactivity towards 99Tc was then probed using a 99mTc(VII) tracer at 0.5 mm which were removed by hand picking. Grains were generally coated with

138

clay-sized iron oxides. X-ray fluorescence confirmed that the sediment comprised Si (30.4 wt

139

%), Al (9.4 wt %), Fe (5.0 wt %), K (3.3 wt %), Mg (1.3 wt %), Na (1.2 wt %), Ti (0.4 wt %), Ca

140

(0.4 wt %), Mn (0.1 wt %), and Sr at 102 ppm. The total iron was approximately 890 mmol kg-1

141

and between 80-100 mmol kg-1 of the sediment Fe(III) was extractable using a 1 hour 0.5 N HCl

142

digestion, an indicator of bioavailable Fe(III)

143

Sellafield sediment was readily bioavailable consistent with other studies on Sellafield near

144

surface materials [14, 15, 31]. The total organic carbon content of the soils was determined as 0.13 ±

145

0.01 % using a LECO CR-412 Carbon Analyser. After equilibration with synthetic groundwater

146

for 24 hours, the sediment pH was 7.2, bracketing typical on site pH values (pH 5 – 8) [12].

[51]

. This suggested ~10 % of the total iron in the

147 148


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Column set up

150

Acrylic columns (20 cm x 3.2 cm i.d.), total volume 160 cm3, were packed with ~ 160 g (volume

151

145 cm3) of moist sediment. A plug of glass wool (~ 1 cm) was used to cap the top and bottom

152

of the columns to minimize sediment movement and ~1 cm quartz sand to optimize fluid flow in

153

the columns (SI Figure 1). The synthetic groundwater was pumped upwards through the columns

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with a flow rate of 4.1 ml hr-1 (5 x 10-6 ms-1) using a Watson–Marlow peristaltic pump. Flow was

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established in columns using filtered (< 0.45 µm) synthetic regional groundwater of composition:

156

KCl (0.09 mmol L-1), MgSO4.7H2O (0.4 mmol L-1), NaNO3 (0.3 mmol L-1), NaCl (0.16 mmol L-

157

1

158

representative of the regional groundwater near Sellafield and in selected column treatments was

159

augmented with 3 mmol L-1 acetate as an electron donor (Table 1). Acetate was selected as an

160

electron donor as it has been used extensively in past work on anaerobic processes including

161

Fe(III) reduction and has been shown to promote the generation of alkalinity and raised pH in

162

Sellafield sediment microcosms with elevated nitrate concentrations which is potentially

163

beneficial in the treatment of strontium-90 in groundwaters

164

concentrations and organics derived from anthropogenic activities are present in selected

165

groundwater monitoring wells on site[12] and the redox state of the subsurface is variable with

166

groundwater oxygen present at up to 0.3 mmol L-1

167

temperature in the dark throughout the experimental programme. The synthetic groundwater was

168

stored in sterile reservoirs at room temperature in the dark and was refreshed every 6 days.

169

Throughout the experiment the pH of the influent groundwater remained at 7.2 and

170

concentrations of acetate, nitrate and sulfate remained within 10 % of the target values in Table

171

1. The initial pore volume of the column prior to experimentation was determined by the addition

), CaCO3 (0.1 mmol L-1), and NaHCO3 (2.8 mmol L-1)[52]. The synthetic groundwater is

[53]

[14,31]

. Indeed, both elevated nitrate

. Columns were maintained at room


8

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of a 5 ppm Br- tracer to each column which was then monitored to define the Br- breakthrough

173

curve (SI Figure 1).

174 175

Column treatments

176

Experiments were separated into “bioreduction” and “post-bioreduction” treatments. For the first

177

60 days, five of the six columns were bioreduced through addition of 3 mmol L-1 sodium acetate,

178

which has been shown to be an effective bio-stimulant for representative Sellafield sediments [14,

179

15, 31]

180

as such represented a ‘natural attenuation’ control (Table 1). Strontium (12 µmol L-1) was added

181

continuously to all six columns. The treated columns and control were run for 60 days, the flow

182

rate stopped and within 4 hours they were then imaged at a gamma camera imaging facility at the

183

Nuclear Medicine Department, Central Manchester University Hospitals. When positioned in the

184

gamma camera, the columns were then spiked with 99mTc (~7 MBq in 1 ml; 3.5 x 10-13 mols) and

185

the flow restarted to image

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radioactive decay period was necessary to allow the 99mTc (half-life 6 hrs) to decay to levels that

187

would allow safe handling; here, the columns were capped and stored without pumping at 4oC in

188

the dark for 5 days. After decay, the columns were transported back to the laboratory where flow

189

was re-instigated under different experimental regimes for a further 50 days. During this post-

190

bioreduction period, a range of different treatments (Table 1) were undertaken to examine 99mTc

191

reactivity: (A) the non-acetate amended groundwater system representing a natural attenuation

192

control; (B) the system with continual acetate amendment; (C) a bioreduced system where

193

acetate was then pulsed (5 days acetate additions, 15 days no acetate additions); (D) a bioreduced

. The remaining column was pumped with synthetic groundwater only (i.e. acetate free), and

99m

Tc behavior under flow conditions. After imaging for 12 hours, a


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system with no further acetate amendment; (E) a bioreduced system where air was then bubbled

195

into the synthetic groundwater (0.31 mM O2); and (F) a bioreduced system where nitrate was

196

then added at elevated levels.

197 198

Sampling and geochemical analysis

199

The column influent and effluent were monitored at regular intervals during experiments. In the

200

effluent, pH and the concentrations of acetate, SO42-, NO2-, NO3-, and Fe(II), were measured to

201

track the progress of terminal electron accepting processes and total Fe, Mn, Al, Ca, Sr and Mg

202

to assess changes in sediment geochemistry. Eh and pH were measured immediately using

203

calibrated electrodes (Denver-Basic). Porewater NO2-, Mn, and Fe(II) were measured

204

spectrophotometrically [54-56]. Acetate, SO42-, NO3- and in tracer tests, Br- were measured by ion

205

chromatography on samples stored at 4 °C prior to analysis. Cation concentrations (total Mg, Al,

206

Fe, Mn, Sr and Ca) were measured by ICP-AES on acidified (2% HNO3) samples. The dissolved

207

O2 concentration of the influent synthetic groundwater was periodically measured using the

208

Winkler titration [57].

209

At experiment end points (115 days), columns were extruded and sampled under a N2

210

atmosphere at 2 cm intervals and bioavailable Fe(II) as a proxy for Fe(III)-reduction was

211

measured. Here, sediment samples were digested in 0.5 N HCl for one hour and aqueous Fe(II)

212

(filtered < 0.2 µm) was measured by ferrozine analysis

213

sulfate reduction with noticeable blackening of the sediment evident. X-ray fluorescence was

214

conducted to ascertain the chemical composition of the sediment (Thermo ARL 9400 XRF).

215

Additionally, at experiment end points the chemical composition of the sediment was measured

[51]

. Sediments showed evidence for


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by microwave digestion of 0.1 g of sediment in 2 ml 50 % HF and 2 ml 16 M HNO3 followed by

217

analysis by ICP-AES. Finally, sequential extractions were used on selected samples to explore

218

differences in the distributions of Fe and Sr using operationally defined, chemical lixivants and

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under anaerobic conditions where appropriate

220

included a 1 hour extraction with 1 M MgCl2 (exchangeable fraction), a 24 hour extraction with

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1 M sodium acetate (carbonate associated fraction), a 24 hour 0.1 M ammonium oxalate

222

extraction (oxidisable fraction), a 6 hour 6 M H2O2 extraction (reducible fraction), and a 12 hour

223

aqua regia extraction (residual fraction).

[58-59]

. The sequential extraction methodology

224 225

Gamma camera imaging and analysis

226

99m

227

gamma emission (140 keV). Imaging with the isotope is common in nuclear medicine, but has

228

also been applied innovatively in environmental systems

229

reactivity of the treated sediments toward Tc was assessed using

230

(bioreduction end-point) and 115 days (post-bioreduction treatment and experiment end-point)

231

columns were transported to the Nuclear Medicine Centre at the Manchester Royal Infirmary for

232

imaging. Here,

233

pertechnetate in deionized water. The imaging was initiated with introduction of a 1 ml spike of

234

~7.0 MBq

235

tubing at the bottom of the column at a flow rate of 5 x 10-6 m s-1. Gamma images were acquired

236

every 20 minutes on a Siemens Symbia T6 dual-headed gamma camera and were processed

237

using GE Xeleris software. This provided a decay corrected vertical 99mTc gamma image through

Tc is a metastable nuclear isomer of

99m

99

Tc with a 6 hour half-life which decays to

[15, 22-24, 50, 60]

99

Tc by

. In the current study the

99m

Tc imaging. At 60 days

99m

TcO4- was introduced from a dilution of pharmacologically pure sodium

Tc (3.5 x 10-11 g

99m

Tc; 3.5 x 10-13 moles; 3.5 x 10-10 mol L-1) into the influent


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the columns. The flow rate (5 x 10-6 m s-1) allowed ~1.2 pore volumes to pass through each

239

column over the 12 hour imaging period during which time the 99mTc was diluted approximately

240

50 fold (from 3.5 x 10-10 mol l-1 to a final concentration in the order of 7 x 10-12 mol l-1). Decay

241

correction was applied after data acquisition and the errors reported on

242

the γ-camera counting error.

99m

Tc activities refer to

243 244

Reactive transport modelling and geochemical modelling

245

The decay corrected radiometric data coupled to the time and flow rate information for the

246

experiments allowed the transport of

247

STANMOD user interface

248

was used to predict the saturation index of key Fe(II) bearing minerals.

[61]

99m

Tc to be modelled using the CXTFIT code with the

. For thermodynamic modelling, PHREEQC-2 (Minteq database)

249 250

Results and discussion

251

Column geochemistry during biostimulation with acetate: days 0-60

252

In the control column pumped with synthetic groundwater not amended with acetate (Column A;

253

which represents natural attenuation conditions), minor depletion of effluent nitrate, and no

254

significant change to Fe, or sulfate concentrations were observed over 60 days (96 pore volumes)

255

(Figure 1). This confirmed that the organic carbon in the sediment (total organic carbon 0.13 ±

256

0.01 %) did not stimulate terminal electron accepting processes beyond nitrate reduction during

257

the experiment. In contrast, bioreduction by continuous addition of 3 mmol L-1 acetate resulted in

258

microbially-mediated metal reducing conditions in all of the other columns (systems B-F) within

259

28 days (Figure 1). Past work using batch experiments shows that Sellafield sediments host a


12

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complex microbial community with a diverse range of bacterial phyla. When stimulated with an

261

electron donor such as acetate, and/or lactate, the diversity of the microbial community was

262

reduced and an increase in close relatives of known nitrate reducing species of the order

263

Bacillales (e.g. Bacillus niacini) and Pseudomonadales (e.g. Pseudomonas peli), and metal

264

reducing species of the orders Clostridiales (e.g. Desulfosporosinus sp.; Alkaliphilus

265

crotonatoxidans; Desulfitobacterium metallireducens), Burkholderiales (e.g. Rhodoferax

266

ferrireducens), Enterobacteriales (e.g. Serratia sp.) and Desulfuromonadales (e.g. Geobacter sp.)

267

were observed

268

acetate concentration in the column effluents between 20 - 60 % of that added (Figure 1B).

269

Complete removal of 0.3 mmol L-1 nitrate, added continuously in the influent synthetic

270

groundwater, occurred within 3 days (5 pore volumes) and gas evolution occurred in the columns

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implying active denitrification to N2 or N2O. Nitrate levels then remained below the detection

272

limit over the first 60 days confirming robust nitrate-reducing conditions had developed within

273

the acetate-amended systems (Figure 1C). Soluble manganese was also detected in the column

274

effluents after 10 days indicating that active Mn(IV) reduction was occurring (data not shown).

275

Aqueous Fe(II) was detected at low concentrations in effluents from 28 days confirming that

276

Fe(III) reduction had developed (Figure 1D). Past research suggests that Fe(II) ingrowth into

277

bioreducing sediments precedes the appearance of Fe(II) in solution, implying that active Fe(III)

278

reduction likely started within column sediments before 28 days

279

the influent synthetic groundwater at 0.4 mmol L-1 but did not significantly decrease in the

280

column effluents over the first 60 days (96 pore volumes) indicating that sulfate-reducing

281

conditions had not developed at the first

282

99m

[7,14,62]

. Consumption of influent acetate in columns B-F was occurring with the

[24]

. Sulfate was also present in

99m

Tc imaging at 60 days (Figure 1E). Thus the first

Tc scan was completed on column systems (B-F) that were predominantly Fe(III)- rather than


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sulfate-reducing. Throughout the experiment, the effluent pH values in acetate amended systems

284

were between 7.5 - 7.8, consistently higher than the pH in the unamended control column

285

measured consistently at pH 7.2. This was presumably a result of alkalinity generated as a result

286

of nitrate reduction (CHCOO- + 4NO3- → 4NO2- + CO2 + H2O; CHCOO- + 2NO2- + 2H+ →

287

2N2O + HCO3- + CO2 + H2O + 2OH-; CHCOO- + 4N2O → 4N2 + HCO3- + CO2 + H2O) and

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Fe(III) reduction (CHCOO- + 8FeOOH + 15H+ → 8Fe2+ + 2HCO3- + 12H2O) [14, 31].

289 290

Imaging Technetium-99 behavior after 60 days of acetate amendment

291

In the unamended oxic control column,

292

in previous work where reducing conditions were not present, sub-nanomolar TcO4- behaved as a

293

conservative tracer

294

good agreement with that of the Br- tracer confirming largely conservative behavior (SI Figure

295

1). The data provided effective porosity estimates of between 41 – 43 % for the columns,

296

measured from five Br- and one

297

column, a best fit model of Br- transport was simulated using STANMOD software (based on

298

CXTFIT code [61]). Here, the model parameters gave a dispersion coefficient of 0.20 cm2 h-1 and

299

a velocity of 4.0 x 10-6 m s-1. The 99mTc transport data under oxic conditions agreed well with the

300

Br- curve, with values of 0.21 cm2 h-1 and 4.2 x 10-6 m s-1 for the dispersion coefficient and

301

velocity respectively (SI Figure 1C). This experimentally derived velocity is consistent with the

302

pumping value of 5 x 10-6 m s-1 and is typical of flow rates through sand and gravel in the

303

Sellafield shallow subsurface of approximately 10-6 m s-1 [12, 15, 63].

[15, 50]

99m

Tc was not retained in the column (Figure 2) and, as

. Indeed, in this column, the breakthrough curves for TcO4- showed

99m

Tc breakthrough curve (SI Figure 1C). For the unamended


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99m

Tc (~ 7 MBq; ~ 3.5 x 10-13

304

Gamma camera imaging after the addition of a 1 ml spike of

305

moles) to each flowing column at 60 days showed consistent results across all acetate amended

306

columns with > 99 % of the added spike retained on contact with the first 6 cm of the column (5

307

cm of sediment excluding the glass wool plug) (Figure 2; SI Table 1). This reactive zone in the

308

columns contained approximately 27 g of sediment and, assuming uniform Tc distribution,

309

yielded retention of Tc concentrations of approximately 1.3 x 10-14 mol g-1 on sediments. This

310

fast retention of

311

exposure to actively Fe(III)-reducing sediment [15, 22-24]. In all systems, technetium concentrations

312

were well below the theoretical solubility limit of hydrous TcO2 (~ 10-8 mol l-1 [64]). This implies

313

sequestration occurred via reduction with surface bound Fe(II) (e.g. TcO4- + 3Fe2+ + 9H2O →

314

TcO2.2H2O(s) + 3Fe(OH)3(s) + 5H+)[3, 9]; with resultant sorption of the reduced Tc [22-24].

99m

Tc in Fe(III)-reducing experiments is consistent with past results from

99m

Tc

315 316

Technetium behaviour in post bioreduction treatments

317

After bioreduction for 60 days and radiological decay storage without flow at 4 °C for 5 days, the

318

experiment continued between 65 – 115 days with several different column treatments. The

319

reactivity of columns to

320

column perturbation treatments were: (A) an non-acetate amended groundwater system; (B) a

321

system that had acetate continuously added to it; (C) a system where acetate addition was pulsed

322

(5 days electron donor addition, 15 days no electron donor addition); (D) a system with no

323

further acetate addition (re-oxidation); (E) an oxygenated system where air was bubbled into the

324

synthetic groundwater prior to addition to the column (0.31 mmol L-1 O2); and (F) a system

325

where 10 mM nitrate was added to explore the impact of high nitrate concentrations on Tc

99m

Tc after these different perturbations was re-tested at 115 days. The


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immobilization after electron donor addition had ceased. These scenarios were chosen to explore

327

a range of credible in-situ delivery options for groundwater treatment at nuclear sites (constant

328

electron donor and staggered electron donor amendments) as well as explore the impact of

329

oxidative perturbations on Tc immobilization in the sediments. After reaction with 99mTc at 115

330

days and imaging, columns were then left to radiologically decay at 4 °C without pumping, and

331

at 120 days, the non-active sediments were extruded, sectioned and then underwent a range of

332

geochemical analyses.

333

In treatment A, the non-acetate amended synthetic groundwater system, no significant

334

geochemical changes were observed between days 65 and 115 (Figure 2). After sediment

335

sectioning, a 0.5 HCl extraction was possible at 2 cm intervals and this showed no detectable

336

Fe(II) was present in the sediments (Figure 2). Furthermore, sequential extractions on the end-

337

point column sediment from column A showed that greater than 90 % of Fe was present in the

338

‘residual’ fraction presumably as refractory clays or other recalcitrant phases (SI Table 2). The

339

spike of

340

conservative tracer as expected in the absence of significant Fe(II)[14, 15, 24].

341

In treatment B, the column continuously stimulated with 3 mmol L-1 acetate for 115 days,

342

immobilization of

343

with greater than 90 % of the Tc retained in the bottom 2 - 4 cm of the column where the 99mTc

344

first contacted the sediment (Figure 2; SI Table 1). Assuming approximately 17 g per 2 cm

345

section of sediment in the column, this represents retention of approximately 2.0 x 10-14 mol g-1

346

99m

347

such that robust sulfate reduction had developed at 65 days after pumping was restarted:

348

CHCOO- + SO42- → 2HCO3- + HS- (Figure 1E). Sulfate was present at 0.4 mmol L-1 in influent

TcO4- added to this system was not retained and the pertechnetate behaved as a

99m

99m

Tc(VII) within the column was essentially complete (greater than 99.9 %)

Tc (Figure 2). Here, bioreduction progressed through nitrate and Fe(III)/Mn(IV) reduction,


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349

synthetic groundwater, and concentrations in the column effluent had dropped to zero by 115

350

days. At the experiment end-point, the sediments were black and smelled of H2S when reacted

351

with 0.5 M HCl suggesting the presence of FeS. The 0.5 N HCl extractable Fe was at

352

approximately 50 % Fe(II)/Fe(III) throughout the column (Figure 2). In this, and all columns

353

during any acetate amendments, minimal aqueous Fe (< 10 µmol L-1) was detected in column

354

effluents, presumably due to the low solubility of Fe(II) at the experimental pH of 7.5 or greater

355

(Figure 1). Indeed, mass balance considerations confirm that less than 0.5 % of the total

356

bioavailable Fe(III) in the column was reductively removed as soluble Fe(II) in effluents (SI

357

Section 1). Furthermore, thermodynamic modelling of the column effluents at 115 days predicted

358

super-saturation with regard to Fe(II)-bearing minerals siderite (FeCO3), magnetite (Fe3O4) and

359

makinawite (FeS) (SI Table 3), supporting experimental observations. Finally, for column B, the

360

sequential extraction data suggested a significant increase in Fe present in the ‘carbonate

361

associated’ fraction throughout the column after bioreduction (SI Table 2). This is consistent

362

with a significant increase in Fe(II)-bearing mineral phases (e.g. siderite (Fe(II)CO3), poorly

363

crystalline monosulfides (FeS), and vivianite (Fe(II)3(PO4)2.8H2O)) extracted during the

364

“carbonate associated” lixivant

365

acetate clearly led to the formation of FeS resulting in sulfidic sediments. Indeed, as observed in

366

previous studies, a rapid rate of 99Tc immobilization was occurring which was comparable to, if

367

not faster than, that observed in actively Fe(III)-reducing sediment where FeS was not present [14,

368

19,22, 67-68]

369

In treatment C, the ‘pulsed’ electron donor column, where 3 mmol L-1 acetate was pumped into

370

the column at the standard flow rate for 5 days out of every 20 days, there were periods of

371

biogeochemical activity stimulated by the electron donor addition. For example, whilst acetate

[65-66]

. As discussed, in column B continuous stimulation with

(Figure 2).


17


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372

addition was occurring, effluent nitrate concentrations decreased (Figure 1C), Fe(II) was

373

detected transiently at low concentrations in solution (Figure 1D), and sulfate concentrations

374

decreased (Figure 1E). By contrast, during periods of no acetate addition, nitrate and sulfate were

375

observed in column effluent indicating that biogeochemical activity had slowed. The 99mTc spike

376

was introduced during a period where no acetate had been added to the column for 15 days, and

377

analysis of the sectioned sediments for 0.5 N HCl extractable Fe(II) showed no Fe(II) in the 0-6

378

cm fractions suggesting significant reoxidation to Fe(III) had occurred, with Fe(II) increasing up

379

to approximately 40 % Fe(II) above 6 cm (Figure 2). Retention of the added Tc occurred

380

between 6 and 18 cm in the column where 0.5 N HCl extractable Fe(II) was present and clearly

381

above the zone of reoxidised sediment that had developed in the first 0–6 cm during 15 days of

382

synthetic groundwater flow without acetate (Figure 2). In terms of retention, 99 % of the

383

spike was distributed across the diffuse zone between 6 - 18 cm (Figure 2; SI Table 1). Despite a

384

vast stoichiometric excess of Fe(II) in sediments above 6 cm in the column (~5 x 10-3 moles of

385

reactive Fe(II) in the 6-18 cm region compared to 10-13 moles of Tc), the sediments were less

386

reactive towards

387

Indeed, the

388

first 4 cm, as in the acetate stimulated experiment at 60 days (Figure 2). This amounts to

389

retention of approximately 3.8 x 10-15 mol g-1 Tc compared to approximately ~2.0 x 10-14 mol g-1

390

Tc in the 60 day acetate stimulated columns. This suggests that oxygen penetration led to

391

reoxidation of the most reactive Fe(II) species, thereby leading to a reduction in reactive Fe(II) in

392

the sediments.

393

In treatment D, the column system where no further acetate was added between days 65 - 115, no

394

further Fe(II) was detected in the column effluent, whilst both nitrate and sulfate were present.

99m

Tc

99m

TcO4- when compared to the Fe(III)-reducing sediments imaged at 60 days.

99m

Tc spike was spread over 12 cm of the column rather than fully retained in the


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395

Sectioning and analysis of this column after 50 days showed that substantial re-oxidation had

396

occurred and Fe(II) was present only in the uppermost sections of the column, between 10–18

397

cm (Figure 2; SI Table 1). As observed in the pulsed acetate column (treatment C), the remaining

398

Fe(II) showed lower reactivity towards

399

region rather than a tight, reactive band occurring. Approximately 80 % of the spike was retained

400

in the 11-18 cm zone on 75 g of sediment giving a concentration of ~ 4.0 x 10-15 mmol g-1 in the

401

partially reoxidised system. In addition, mass balance calculations showed ~ 20 % of the spike

402

had been transported out of the column and was not retained. The significant re-oxidation in the

403

sediment was presumably due to the oxygen in the influent synthetic groundwater that was

404

measured at ~ 0.2 mmol L-1 O2. This was consistent with expected dissolved oxygen saturation

405

levels at 20 °C, and was sufficient to reoxidise approximately 75 % of bioreduced Fe(II) in the

406

column after movement of 80 pore volumes of oxygenated water through the column.

407

In treatment E, the system where oxygenated synthetic groundwater was pumped into the column

408

from 65-115 days,

409

similar to that of the oxic column (Figure 2; SI Table 1). Here, constant aeration of the influent

410

synthetic groundwater resulted in a dissolved oxygen concentration measured at 0.31 mmol L-1

411

O2 (equivalent to dissolved oxygen saturation at 15 °C). Indeed, calculations showed that

412

complete reoxidation of the residual Fe(II) species was expected to occur after transport of 80

413

pore volumes of O2 saturated water through the columns. This was confirmed by analysis of the

414

sectioned core which showed the 0.5 N HCl extractable Fe in sediments after reaction was

415

present as Fe(III) with no detectable Fe(II). This presumably explained the lack of reactivity to

416

TcO4-(aq).

99m

Tc with a diffuse zone of retention in the 11-18 cm

99m

Tc was not retained on sediments and the Tc reactivity showed behavior


19


Page 20 of 40

417

Finally, in treatment F, elevated nitrate (10 mmol L-1) was added to the synthetic groundwater

418

(without added acetate) to explore the impact of high nitrate concentrations in groundwaters.

419

Here, it was hypothesized that column reoxidation may be expected to be accelerated by the

420

microbial reduction of nitrate coupled to the oxidation of Fe(II)

421

shown that for U(IV), intermediate species such as NO2- and Fe(III) from nitrate-dependent

422

microbial Fe(II) oxidation are key controls on reoxidation to U(VI) [3, 69]. However, in the current

423

study nitrate proved to be unreactive in the column and no transient nitrite was detected in

424

effluents. Overall, Column F behaved in a comparable way to treatment D, the low nitrate system

425

with a reoxidised zone developing rather slowly in the column. Indeed, analysis of the sectioned

426

column showed bioavailable Fe(II) was present in the upper 6 cm of the column (similar to the

427

low nitrate system). The similarity between Columns D and F, and the lack of nitrate utilisation

428

in these systems suggested that that elevated nitrate had no significant effect on Fe(II) oxidation

429

rates in these experiments and that reoxidation via O2 was the dominant process. Approximately

430

75 % of

431

behaviour to column D. It is possible that the presence of even low O2 in the influent

432

groundwater would inhibit nitrate-dependent Fe(II) oxidation, as many microorganisms

433

identified as able to couple Fe(II) oxidation to nitrate reduction do so under strict anaerobic

434

conditions [70].

[42, 63]

. Indeed, past work has

99m

Tc spike was retained within the upper 6 cm of column F, confirming comparable

435 436

Behaviour of strontium in column systems

437

Strontium was introduced continually into all columns in the synthetic groundwater at 12 µmol

438

L-1 and its retention was monitored by analysis of Sr2+ in column effluents (Figure 1F) and at the

439

experiment end-point by total acid digestion of the sediment followed by ICP-AES analysis. The


20

Page 21 of 40


440

pH in the oxic control remained steady at pH 7.2 throughout the experiment, and in other

441

columns was between 7.5 - 7.8 (Figure 1A). Effluent concentrations of Sr2+ trended upwards over

442

the duration of the experiment in all systems presumably as sediment Sr2+ sorption sites became

443

saturated, and with column B effluents showing the most significant Sr2+ retention over 115 days

444

(Figure 1F). Furthermore, experimental end point measurements of Sr in sediments by total acid

445

digestion (SI Table 4) showed enhanced Sr retention in column B (continuous acetate

446

amendment, 149 mg kg-1) compared to both the other columns and the background Sr

447

concentration of 102 mg kg-1 measured in the sediments. Here, the pH of the effluent was

448

amongst the highest of the column treatments (pH 7.7) and there was clear evidence for the

449

robust development of sulfate reducing conditions by the end of the experiment (Figure 1A, 1E).

450

Thermodynamic modelling of this system using an Eh of -240 mV (representative of typical

451

sulfate-reducing conditions), showed several Fe(II)-bearing mineral phases (siderite, magnetite,

452

and mackinawite) were modestly oversaturated, but undersaturation was predicted for SrCO3 (SI

453

Table 3). It is therefore probable that enhanced Sr2+ retention in this system was dominated by

454

sorption to either mineral phases and / or microbes [41-43, 71-72]. Finally, it should also be noted that

455

end point sediment total acid digests for columns A and C – F were all broadly consistent (91 ± 4

456

mg kg-1) with no Sr2+ enhancement above the initial concentration measured by XRF (102 mg

457

kg-1). Minor loss of Fe from the bioreducing columns was measurable in the effluents, most

458

notably when the columns had been standing for 5 days after initial acetate treatments (Figure

459

1D) and presumably due to reductive dissolution of Fe(III) minerals. However, this loss did not

460

appear to significantly affect Sr2+ sorption to the sediments (Figure 1F). Indeed, significant

461

biogeochemical perturbations involving different reduction and reoxidation cycles, appeared to


21


Page 22 of 40

462

have very little impact on the total Sr2+ retention over the four-month experiment and similar to

463

past work on static microcosm experiments using Sellafield sediments. [31]

464 465

Environmental Implications

466

Overall, these findings are important for understanding contaminant migration at nuclear “mega

467

sites” and the potential for in-situ bioreduction treatment of radionuclide releases to the

468

subsurface. In the absence of an added electron donor, oxic unstimulated sediment had a high

469

capacity for Sr sorption but was unreactive towards

470

reducing conditions can be achieved with the addition of 3 mmol L-1 acetate for less than 60 days

471

and that

472

demonstrates the potential for co-treatment of

473

where acetate was added constantly for 115 days (treatment B), increased Sr sorption capacity

474

was observed compared to the unamended control column. Whilst small-scale column studies do

475

not provide a substitute for full scale field testing,

476

dynamic sediment columns to be tested non-destructively after bioreduction and in post-

477

bioreduction treatments in complex heterogeneous and biogeochemically evolving sediment

478

columns. Visualizing

479

(treatment C), sediments became less reactive between pulses and enhanced Tc migration was

480

observed. When electron donor addition ceased (treatment D and E), reoxidation was significant

481

(> 75 % over 50 days), depending on the oxygen concentration of influent synthetic

482

groundwater. Here,

483

reoxidised sediments (treatment D). However, increased nitrate concentrations (10 mmol L-1,

99

Tc. Bioreduction experiments show that

99m

Tc was retained with no measurable loss in Sr sorption capacity. This clearly 99

Tc and

90

Sr via in-situ bioreduction. Moreover,

99m

Tc imaging allowed the reactivity of

99m

Tc transport in columns showed that when electron donor was pulsed

99m

Tc was only retained when Fe(II) was still partially present in the


22

Page 23 of 40


484

representative of some contaminant plumes at nuclear facilities) had little or no effect on

485

reoxidation rates in these experiments (treatment F). Indeed, high nitrate columns contained the

486

same proportion of Fe(II) after 50 days as low nitrate columns. Results highlight the ability of

487

carefully established column experiments to provide essential information which is crucial in

488

bridging the gap between laboratory batch experiments and field-scale testing and allow more

489

complex and potentially larger scale test systems to be developed. This information is directly

490

pertinent to the Sellafield nuclear facility in contaminated land management planning, and has

491

implications for other nuclear legacy sites where Tc and Sr are found as co-contaminants in

492

groundwater.

493 494


23


495

Page 24 of 40

ASSOCIATED CONTENT

496

Supporting Information. Tables and Figures showing additional geochemical and imaging

497

results are available free of charge via the Internet at http://pubs.acs.org.

498 499

Acknowledgements

500

This work was co-funded by Sellafield Ltd. and The University of Manchester EPSRC Impact

501

Acceleration Account. We thank Alastair Bewsher and Paul Lythgoe for help in data acquisition

502

and James Graham, National Nuclear Laboratories for helpful discussions on site groundwater

503

conditions.

504 505


24

Page 25 of 40

506 507


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Talanta. 1997, 24, 752-753; DOI 10.1016/0039-9140(77)80206-3.


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55. Viollier, E.; Inglett, P.W.; Hunter, K.; Roychoudhury, A.N.; Van Cappellen, P. The

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ferrozine method revisited: Fe(II)/Fe(III) determination in natural waters. Appl. Geochem.

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56. Harris, S.J.; Mortimer, R.J.G. Determination of nitrate in small water samples by the

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cadmium-copper reduction method: A manual technique with application to the

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interstitial waters of marine sediments. Inter. J. Environ. Anal. Chem. 2002, 82, 369-376;

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57. Winkler, L.W. Berichte der Deutschen Chemischen Gesellechaft, 1888, 21, 2843.

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58. Tessier, A.; Campbell, P.G.C.; Bisson, M. Sequential extraction procedure for the

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speciation of particulate trace metals. Anal. Chem. 1979, 51, 884-851; DOI

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59. Keith-Roach, M.J.; Morris, K.; Dahlgaard, H. An investigation into technetium binding in

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sediments. Marine Chem. 2003, 81, 149-162; DOI 10.1016/S0304-4203(03)00014-8.

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60. Aarkrog, A.; Carlsson, L.; Chen, Q.J.; Dahlgaard, H.; Holm, E.; Huynh-Ngoc, L.; Jensen,

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L.H.; Nielsen, S.P.; Nies, H. Origin of technetium-99 and its use as a marine tracer.

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Nature. 1988, 335, 338-340; DOI 10.1038/335338a0. 61. Simunek, J.; van Geunchten, M. Th.; Sejna, M.; Toride, N.; Leij, F.J. The STANMOD

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Computer software for evaluating solute transport in porous media using analytical

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solutions of convection-dispersion equation. U.S. Salinity laboratory agricultural research

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service. U.S. Department of Agriculture, Riverside, California, 1999.

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62. Newsome, L; Morris, K; Trivedi, D; Atherton, N; Lloyd, J.R. Microbial reduction of

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uranium(VI) in sediments of different lithologies collected from Sellafield. Appl.

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Geochem. 2014. 51, 55–64; DOI 10.1016/j.apgeochem.2014.09.008.

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63. Hubbard, C. G.; West, L. J.; Morris, K.; Kulessa, B.; Brookshaw, D.; Lloyd, J. R.; Shaw,

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S., In search of experimental evidence for the biogeobattery. J. Geophys. Res.: Biogeosci.

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64. Hess, N.J.; Xia, Y.X.; Rai, D.; Conradson, S.D. Thermodynamic model for the solubility

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of TcO2.xH2O(am) in the aqueous Tc(IV)-Na+-Cl−-H+-OH−- H2O system. J. Sol. Chem.

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2004, 33,199-226; DOI 10.1023/B:JOSL.0000030285.11512.1f.


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investigation of two sequential extraction techniques applied to anaerobic canal bed mud.

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Environ. Geochem. Health. 2000, 22, 281–296; DOI 10.1023/A:1006743630918.

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66. Torres, E.; Auleda, M. A sequential extraction procedure for sediments affected by acid

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mine drainage. J. Geochem. Explor. 2013. 128, 35-41; DOI

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10.1016/j.gexplo.2013.01.012.

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67. Liu, Y.; Terry, J.; Jurisson, S. Pertechnetate immobilization with amorphous iron sulfide. Radiochim. Acta. 2008, 96, 823–833; DOI 10.1524/ract.2008.1528. 68. Lee, J.; Zachara, J.M.; Fredrickson, J.K.; Heald, S.M.; McKinley, J.P.; Plymale, A.E.;

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Resch, C.T.; Moore, D.A. Fe(II)- and sulfide-facilitated reduction of 99Tc(VII)O4- in

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microbially reduced hyporheic zone sediments. Geochim. Cosmochim. Acta. 2014, 136,

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247-264; DOI 10.1016/j.gca.2013.08.017.

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69. Elias, D. A.; Krumholz, L. R.; Wong, D.; Long, P. E.; McKinley, J. P.; Suflita, J. M.

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Characterization of microbial activities and U reduction in a shallow aquifer

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contaminated by uranium mill tailings. J. Microb. Ecol. 2003, 46, 83-91; DOI

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70. Carlson, H.K.; Iain, C.C; Melnyk, R.A.; Coates, J.D. Toward a mechanistic

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71. Faison, B.S.; Cancel, C.A.; Lewis, S.N.; Adler, H.I. Binding of dissolved strontium by

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72. Schultze-Lam, S.; Beveridge, T.J. Nucleation of celestite and strontianite on a

738

cyanobacterial S-layer. Appl. Environ. Microbiol. 1994, 60 (2), 447–453.

739


32

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740 741

Figure 1. Composition of column effluents over 115 days (data collection was paused between

742

60–65 days to allow radiological decay of 99mTc before experiments could restart). Graphs show

743

A) pH, B) acetate, C) nitrate, D) Fe(II), E) sulfate, and F) strontium values and/or concentrations

744

in effluents. Influent acetate was 3 mmol L-1 for columns B-F for days 0-60. Further acetate was

745

continually added at 3 mmol L-1 to column B between 65-115 days, and in pulsed amendments to

746

column C between 80-85 days and 100-105 days. Influent nitrate was 10 mmol L-1 for column F


33


Page 34 of 40

747

from 65-115 days and was added continually at 0.3 mmol L-1 for all other columns. Influent

748

sulfate was 0.4 mmol L-1 throughout the experiment for all columns.


34

Page 35 of 40


749


35


Page 36 of 40

750

Figure 2: Panel I: Gamma camera images of columns at 60 days. Images are at end-point

751

samples (12 hours after introduction of

752

synthetic groundwater had passed through the column. The images show: (A)99mTc spike at the

753

uppermost sections of the unamended control column; and (B-F) 99mTc immobilized in the lower

754

2 cm of sediment in the columns w hich had been amended with 3 mmol L-1 acetate for 60 days

755

(Note: the lowest 2cm section of the column consisted of glass wool and quartz sand packing to

756

improve flow dynamics the next 2 – 4 cm section contained the bottom of the sediment column).

757

Panel II: 99mTc images at the 115 day time point. Images are at end-point samples (10 hours after

758

the introduction of the

759

had passed through the columns which were: (A) the non-acetate amended groundwater system

760

representing a natural attenuation control; (B) the system with continual acetate additions; (C) a

761

bioreduced system where acetate was then pulsed (15 days no acetate additions, 5 days acetate

762

additions); (D) a bioreduced system with no further acetate additions; (E) a bioreduced system

763

where air was then bubbled into the synthetic groundwater (0.31 mM O2); and (F) a bioreduced

764

system where nitrate was then added at elevated levels. The data are co-plotted with the

765

percentage distribution of the

766

sections and the % 0.5 N HCl extractable Fe present as Fe(II) (blue triangles) plotted at the

767

bottom of the 2 cm sections sampled.

99m

Tc) and after approximately 1.2 pore volumes of

99m

Tc) and after approximately 1 pore volumes of synthetic groundwater

99m

Tc spike (black squares) calculated for the mid-point of 2 cm

768


36

Page 37 of 40

769


Table 1. Experimental details. Column

Continuous Sr2+ addition

A

12 µmol L-1

B

Input 0-60 days

61-65 days

Input 65-115 days

No acetate amendment

99m

No acetate amendment

12 µmol L-1

3 mmol L-1 acetate

99m

3 mmol L-1 acetate

C

12 µmol L-1

3 mmol L-1 acetate

99m

3 mmol L-1 acetate in 5 day pulses every 20 days

D

12 µmol L-1

3 mmol L-1 acetate

99m

No further amendment, 0.3 mmol L-1 nitrate, 0.25 mmol L-1 oxygen

E

12 µmol L-1

3 mmol L-1 acetate

99m

No further amendment, 0.3 mmol L-1 nitrate, oxygen purged influent (0.31 mmol L-1)

F

12 µmol L-1

3 mmol L-1 acetate

99m

No further amendment, 10 mmol L-1 nitrate, 0.25 mmol L-1 oxygen

Tc decay Tc decay Tc decay Tc decay Tc decay Tc decay

770 771 772


37


Figure 1 254x239mm (96 x 96 DPI)


Page 38 of 40

Page 39 of 40


Figure 2 190x275mm (96 x 96 DPI)



254x190mm (96 x 96 DPI)


Page 40 of 40

Quantifying Technetium and Strontium Bioremediation Potential in

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