Biogeochemical Controls of Uranium ... - ACS Publications

Subscriber access provided by the University of Exeter

Article

Biogeochemical controls of uranium bioavailability from the dissolved phase in natural freshwaters Marie Noele Croteau, Christopher C. Fuller, Daniel J. Cain, Kate Campbell, and George R. Aiken Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02406 • Publication Date (Web): 06 Jul 2016 Downloaded from http://pubs.acs.org on July 13, 2016

Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a free service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are accessible to all readers and citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.

Environmental Science & Technology is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.

Page 1 of 32

Environmental Science & Technology

1

Biogeochemical controls of uranium bioavailability

2

from the dissolved phase in natural freshwaters

3

Marie-Noële Croteau1*, Christopher C. Fuller1, Daniel J. Cain1, Kate Campbell2 and George

4

Aiken2

5

1

U.S. Geological Survey, 345 Middlefield Rd, Menlo Park, CA94025, United States

6

2

U.S. Geological Survey, 3215 Marine St Suite E-127, Boulder, CO80303, United States

7

8

To gain insights into the risks associated with uranium (U) mining and processing, we

9

investigated the biogeochemical controls of U bioavailability in the model freshwater species

10

Lymnaea stagnalis (Gastropoda). Bioavailability of dissolved U(VI) was characterized in

11

controlled laboratory experiments over a range of water hardness, pH, and in the presence of

12

complexing ligands in the form of dissolved natural organic matter (DOM). Results show that

13

dissolved U is bioavailable under all the geochemical conditions tested. Uranium uptake rates

14

follow first order kinetics over a range encompassing most environmental concentrations.

15

Uranium uptake rates in L. stagnalis ultimately demonstrate saturation uptake kinetics when

16

exposure concentrations exceed 100 nM, suggesting uptake via a finite number of carriers or ion

17

channels. The lack of a relationship between U uptake rate constants and Ca uptake rates suggest

18

that U does not exclusively use Ca membrane transporters. In general, U bioavailability

ACS Paragon Plus Environment

1


Page 2 of 32

19

decreases with increasing pH, increasing Ca and Mg concentrations, and when DOM is present.

20

Competing ions did not affect U uptake rates. Speciation modeling that includes formation

21

constants for U ternary complexes reveals that the aqueous concentration of dicarbonato U

22

species (UO2(CO3)2-2) best predicts U bioavailability to L. stagnalis, challenging the free-ion

23

activity model postulate.

24

INTRODUCTION

25

Growing world-wide demand for uranium (U) as an energy source has raised concerns about

26

the human and ecological risks of U extraction and processing. Uncertainties in risks of U

27

mining prompted a 20-year withdrawal of Federal lands near Grand Canyon National Park from

28

future U mining to allow further study, including risks to Federal and State listed threatened and

29

endangered species1. Determining the environmental health effects of U requires understanding

30

the exposure pathways and bioaccumulation mechanisms of this element to aquatic and

31

terrestrial organisms. Although a great deal of information exists on the toxicity of U to

32

freshwater organisms (e.g. van Dam et al.2, and references therein), far less is known about the

33

underlying mechanisms governing U uptake and loss in invertebrates. Uranium is unique among

34

metals in that it induces both chemical and radiological toxicity. However, the risks to the

35

environment from uranium’s chemical toxicity generally surpass those of its radiological

36

toxicity3.

37

Although uranium has no known essential biological role, U is taken up by organisms4-5. The

38

transport mechanisms of U into cells remain largely unknown, however. Possible mechanisms

39

include uptake into cells due to increased membrane permeability likely caused by U toxicity6,

40

uptake through membrane carriers dedicated to an essential element like Ca7 or Fe8, or by


2

Page 3 of 32


41

passive uptake9 (e.g., facilitated diffusion). Anionic species of U like carbonato U complexes

42

(e.g., UO2(CO3)2-2) might also cross membranes via anionic channels as analogues for phosphate

43

or bicarbonate, as anion channels are often large and nonselective10.

44

In contrast, the influence of water chemistry on the bioavailability and toxicity of U is well

45

documented. In general, increase in pH, alkalinity, water hardness, and DOM attenuate U

46

toxicity, either through competition for surface-binding sites (e.g., H+, Ca2+)11-12 or through

47

aqueous complexation13. The free uranyl ion (UO2+2) has been reported to best predict biological

48

responses resulting from dissolved U exposures, in support of the free-ion activity model

49

(FIAM)14. For instance, Fortin et al.15 reported that U uptake by a green alga correlated with the

50

UO2+2 concentration. However, exceptions to the FIAM exist. For instance, concentrations of

51

both UO2+2 and UO2OH+ have been used to predict adverse behavioral responses in a freshwater

52

bivalve16. Because ternary calcium and magnesium uranyl carbonate complexes have been

53

largely omitted from studies relating U speciation to bioavailability (e.g., references 12,15,16),

54

additional exceptions to the FIAM are likely. Formation constants for these ternary complexes

55

have only been recently published17 and used mostly in groundwater modeling studies18. These

56

ternary U complexes can bind the majority of U when Ca concentrations exceed 50 µM, which is

57

the case of most natural freshwaters19. A recent study linked the predominance of these

58

complexes to the low U toxicity observed in humans chronically exposed to U20, suggesting a

59

crucial role for these complexes for bioavailability. Finally, binding of U by DOM is usually not

60

accounted for by many speciation models, resulting in an overestimation of inorganic U

61

species21.

62

In this study, controlled laboratory experiments were conducted to characterize the underlying

63

mechanisms controlling U(VI) bioaccumulation after waterborne exposure in a model species,


3


Page 4 of 32

64

the freshwater snail Lymnaea stagnalis. This herbivore species represents an initial, key step in

65

the trophic transfer of contaminants through aquatic food webs22-23. Specifically, the effects of

66

water chemistry and competing ligands on dissolved U speciation were evaluated using

67

PHREEQC modelling. Formation constants for ternary U complexes and for the binding of U

68

onto DOM were included in the speciation calculations prior to identifying the U species that

69

best predict bioavailability. Experiments were conducted in synthetic freshwaters over a range of

70

pH (from 6 to 8), Ca concentrations (0.4 to 4.2 mM) and alkalinity (0.13 to 1.15 mM), as well as

71

in the presence of a well characterized DOM. Test water also included an artificial stream water

72

(ASW) designed to simulate geochemical conditions relevant to Kanab Creek, a stream in a

73

watershed that drains a portion of the withdrawal area near Grand Canyon National Park1,24.

74

METHODS

75

Experimental organisms. Snails (L. stagnalis) were reared in the laboratory in moderately

76

hard (MOD) water25 (Table 1). They were progressively acclimated to the experimental

77

conditions in 1-L jars for up to 1 week (except for experiment 4, see below). Lettuce was fed to

78

snails during the acclimation period, but withheld during the exposure. Snails of a restricted size

79

were used for each experiment to minimize the potential confounding influence of size on the

80

aqueous U uptake kinetics26 (Table S1).

81

Waterborne exposures (Experiments 1-3). The effect of U speciation on bioavailability was

82

tested at 13.5 °C by varying the chemical composition of the experimental waters (Experiment

83

set 1), by varying the pH (Experiment set 2), as well as by adding complexing organic ligands in

84

the form of a DOM isolate (Experiment 3) (Table 1, Table S2). Experimental systems were well

85

oxygenated and all dissolved U was assumed to be in the +6 oxidation state, U(VI), and referred


4

Page 5 of 32


86

to hereafter as U. A stock solution of U (0.01 N) was prepared by dissolving 1.001 g of

87

UO2(NO3)2·6 H2O in 200 mL of high purity water (18 MΩ cm). pH was measured before and

88

after exposures using an Orion 520A meter and Ross combination electrode calibrated at 13.5

89

°C, and pre-equilibrated in experimental media without U.

90

Calcium concentrations varied 30-fold in the first series of experiments (i.e., from 0.044 to

91

1.45 mM, Table 1). All dissolved Ca species are referred to hereafter as Ca, unless specified. The

92

total dissolved U concentrations in the first series of experiments ranged from < 0.008 to 5260

93

nM (Table S2). These experiments were conducted open to the atmosphere. However respiration

94

increased the pCO2 during exposure, resulting in a decrease in pH of 0.4-0.6 pH units (Table S2).

95

In the second set of experiments, exposures were conducted at a fixed pH (i.e., 6.0, 6.5, 7.0, 7.5

96

and 8.0) in MOD water spiked with U to achieve dissolved U concentrations ranging from < 1 to

97

46 nM. The pH was controlled by imposing an elevated pCO2 by continuously purging (~200

98

mL/min) the exposure media with air using MKS mass flow controllers to mix CO2 with air.

99

Prior to adding the experimental snails, the pCO2 of the gas mixture was adjusted to yield the

100

target pH (Table S2). Experimental pH varied ± 0.05 units during exposure and among U

101

concentrations. The variations in pH and pCO2 in our experiments likely had little effect on L.

102

stagnalis’ acid-base balance and ion (Na+ and Cl-)27, and presumably on U uptake kinetics. In

103

experiment 3, MOD water was amended with a hydrophobic organic acid (HPoA) DOM isolate

104

of a relatively high aromaticity (specific ultraviolet absorbance of 5.4) obtained from the St

105

Louis River (MN) according to the methods of Aiken et al.28 and dialyzed in MOD water using a

106

1000 Da dialysis bag (Suprapor)21. Additional DOM properties can be found in the supplemental

107

information (Table S3). Solutions were made to achieve a dissolved organic carbon


5


108

concentration of 14 mg C L-1. pH was maintained at 7.5 with fixed pCO2 and dissolved U

109

concentrations ranged from 3 to 112 nM.

110

Page 6 of 32

Short exposures (24 h) were employed to minimize the influence of efflux, allowing

111

measurement of gross net influx rates29. Specifically, 10 acclimated snails were transferred to

112

acid-washed 1 L HDPE containers filled with experimental waters spiked with a wide range of

113

dissolved U concentrations. The lowest U exposures were environmentally relevant, i.e., < 10 µg

114

L-1 (19). After exposure, snails were removed from the experimental media, rinsed with ultrapure

115

water, and frozen. Before and after exposure, water samples were taken from each vial. Samples

116

were filtered through a 0.45 µm Millex-HV filter (after exposure only) and acidified with

117

double-distilled 16 N nitric acid (1% final concentration).

118

Ca uptake rates (Experiment 4). To evaluate the influence of Ca on U uptake rates, Ca influx

119

rates were characterized in short-term exposures at a low dissolved U concentration (25 nM).

120

Prior to exposure, snails were progressively acclimated over a 7 day period to synthetic

121

freshwater of varying hardness. Five calcium concentrations were tested, spanning the range of

122

the calcium concentrations of the very soft, soft and moderately hard water (Table S4)25.

123

Isotopically enriched 44Ca (Trace Sciences International, 99.20%) as CaCO3 (dissolved in a

124

stoichiometric equivalent of sulfuric acid) was used in place of CaSO4·2 H2O in the formulation

125

of the synthetic freshwaters.

126

Ten snails were transferred to acid-washed 200 mL glass beakers filled with media spiked with

127

U. Snails were exposed to 44Ca and U for 4 h, and thereafter processed as described above.

128

Before and after exposure, pH was measured and water samples were taken from each vial, as

129

described above (Table S4).


6

Page 7 of 32


130

Elimination (Experiment 5). To determine how strongly U was retained by the snails after

131

waterborne uptake and to quantify the unidirectional U efflux rate constant, we exposed 90 snails

132

to 100 nM of U in the Kanab Creek ASW (Table 1) for 2 d. Snails were not fed during the

133

exposure period. After exposure, snails were removed from the exposure media, rinsed with

134

ASW and distributed into seven 150 mL acid-washed vials partially submerged in a 40-L glass

135

tank filled with ASW. At each sampling time (i.e., 0, 1, 2, 3, 5, 7, 10 and 14 d), snails were

136

collected from an entire vial, rinsed with ultrapure water and frozen. Water was sampled at each

137

sampling time (as described above). Uncontaminated food (lettuce) was provided throughout

138

depuration. Fecal material was removed from each remaining depuration chamber prior to

139

adding fresh food on day 5 to minimize the confounding influence of U recycling.

140

Sample preparation and analysis. To minimize inadvertent metal contamination, labware,

141

vials, and Teflon sheeting were soaked for at least 24 h in acid (15% nitric and 5% hydrochloric),

142

rinsed several times in ultrapure water and dried under a laminar flow hood prior to use.

143

Partially thawed L. stagnalis were dissected using stainless tweezers to remove soft tissues,

144

placed individually on a piece of acid-washed Teflon sheeting and allowed to dry at 40 °C for 3

145

d. Dry weights for soft tissues were determined to the nearest µg on a microbalance (Sartorius

146

Model M2P). Snail tissues were digested in PTFE vials with 16 N double-distilled nitric acid for

147

3 h at 125 °C in an autoclave. Digested samples were diluted with ultrapure water (2% HNO3)

148

and an internal standard (thallium) was added to control signal drift. Diluted samples were

149

filtered (0.45 µm, Pall), and analyzed for U by inductively coupled plasma-mass spectrometry

150

(ICP-MS, PerkinElmer NexION 300Q). The method detection limit (MDL) for U was 2 ng L-1.

151

We also reanalyzed one of our standards after every 10 samples. Deviations from the standard

152

values were < 5% for the analyzed U isotope (238U) at all times.


7


153

Page 8 of 32

Derivation of provisional uptake rate constants. The uptake rate constant from solution, kuw

154

in l g-1 d-1, was used to infer U bioavailability from water. kuw was determined from the slope of

155

the linear regression between U influx into the snail’s soft tissues and the total U exposure

156

concentrations ([U]water) (data from the linear portion of the curve, controls excluded). Uranium

157

elimination was modeled by nonlinear regression using a two-compartment model, as described

158

in the Supporting Information.

159

Speciation calculations. Aqueous U speciation was calculated with PHREEQC30 at 25 °C using

160

the major ion chemistry from Table 1 and Table S2, along with the wateq4 database with

161

aqueous uranium species stability constants from Guillaumont et al.31 and included ternary

162

(Ca,Mg)-U(VI)-CO3 complexes17,32. Thermodynamic data for adjusting many of U stability

163

constants to the experimental temperature are not available. Although the stability constants for

164

the ternary complexes are not part of the most recent compilation of critically evaluated

165

constants31, their documented effect on U solubility, surface complexation and toxicity18,20,33-37

166

warrants inclusion in characterizing effects of U speciation on bioavailability. The pH measured

167

at the end of the experiment and alkalinity, as determined by Gran titration, were held constant in

168

speciation calculations. The resulting calculated pCO2 was consistent with the measured pCO2 of

169

the gas mixer outflow (Table S2). The speciation of U in the systems with DOM was calculated

170

using conditional U-DOM formation constants derived by equilibrium dialysis ligand exchange

171

(EDLE) methodology21 (see below).

172

Determination of U-DOM conditional binding constants using EDLE methods.

173

Conditional binding constants were measured using a modified EDLE method based on van

174

Loon et al.38. Briefly, 5.5 mL of a DOM-containing MOD solution contained within a 1000 Da

175

dialysis bag (Suprapor) was equilibrated with 30 mL of a U-containing MOD solution. The


8

Page 9 of 32


176

DOM-containing (“inner”) solution was made from the same pre-dialyzed DOM isolate at the

177

same concentration (14 mg L-1) used in Experiment 3. The U-containing (“outer”) solutions were

178

made from dilution of a 0.01 M U(VI) stock solution to 3.9 nM, 15.3 nM, 34.3 nM, or 159 nM U

179

and 250 µM EDTA as a competitive ligand. The MOD water used to make the DOM inner and

180

the U outer solutions was equilibrated with CO2 gas to adjust the pH to 7.5 at 13 °C; inner and

181

outer solutions were confirmed to be pH 7.5 ± 0.1 prior to the start of the experiment. Outer

182

solutions were distributed into a 40 mL acid-washed, baked amber vial. The DOM inner solution

183

was sealed inside a dialysis bag and added to the vial containing the outer solution and allowed

184

to equilibrate at 13 °C for 46 hours on an orbital shaker table while the outer solution was

185

bubbled with the correct gas mixture to maintain pH 7.5, which was verified at the end of the

186

experiment. Each experimental condition was performed in triplicate. The conditional binding

187

constant (KDOMU) was calculated as described in the Supporting Information.

188

Results

189

Waterborne uranium uptake (experiments 1-3). Uranium accumulation in snails exposed to

190

dissolved U concentrations > 1 nM was significant compared to controls after 24 h exposure (t-

191

tests, P < 0.01). The snail U background concentration was 0.2 ± 0.08 nmole g-1 (n = 39).

192

Uranium uptake rates were significantly correlated with size in 5 of the 24 treatments in

193

experiment 1 (P < 0.05). The influence of size is expressed as higher variability in U uptake rates

194

(Figure 1). Uptake rates of U into L. stagnalis increased linearly with U concentrations

195

encompassing aqueous U concentrations measured from U mining impacted surface waters39

196

(Figure 1). Uranium influx rates began to go non-linear at U concentrations greater than 100 nM

197

(Figure 1; Figure S2). The shape of the curves was typical of one site saturation kinetics

198

(equation S4, Table S7), suggesting uptake via a finite number of carriers or ion channels.


9


Page 10 of 32

199

Uranium bioavailability was inversely dependent on the Ca and Mg concentrations. The mean

200

values for the U uptake rates into L. stagnalis tended to range from high in the soft water to low

201

in the harder water when dissolved U exposure concentrations ranged from 10 to 1000 nM

202

(Figure 1). As shown in Figure 2, kuw decreased 5-fold over the different exposure waters.

203

Specifically, kuw (± SE, in l g-1 d-1) was 1.6 ± 0.01 in very soft water and declined nearly 3-fold in

204

soft water to 0.60 ± 0.01. kuw was slightly lower in MOD water (0.55 ± 0.02) than in soft water,

205

and decreased further in the synthetic water simulating the inorganic composition of Kanab

206

Creek (0.28 ± 0.01).

207

Uranium bioavailability was also influenced by pH (experiment set 2). In MOD water, kuw

208

increased 1.7-fold when pH increased from 6.0 to 6.5 (Figure 3). Further increases in pH yielded

209

decreases in kuw. Specifically, kuw (± SE, in l g-1 d-1) was highest at pH 6.5 (1.1 ± 0.04) and

210

progressively declined to 0.42 ± 0.02 at pH 8.0. At pH 7.5, kuw decreased six-fold (from 0.39 to

211

0.069 l g-1 d-1) in the presence of DOM (experiment 3), which is consistent with strong aqueous

212

complexation of U(VI) by DOM (star, Figure 3).

213

Calcium uptake (experiment 4). Calcium uptake rates into L. stagnalis increased with

214

increasing Ca concentrations, but Ca influxes rapidly plateaued when Ca concentrations

215

exceeded 100 µM (Figure 4a). Nonlinear regression fit of the mean values to a one site ligand

216

binding saturation model (Equation S4, Supporting Information) yielded a best-fit solution with

217

R2 of 0.56, a half saturation constant (Kd ± SE) of 35 ± 30 µM and a maximum uptake rate (Vmax

218

± SE) of 217 ± 35 µmol g-1 d-1. In contrast, provisional U uptake rate constants estimated by

219

dividing the U influx by the initial U exposure concentration, decreased with increasing Ca

220

concentrations (Figure 4b), except at a Ca concentration of ~80 µM where kuw was greater than at

221

40 µM.


10

Page 11 of 32


222

Speciation. The calculated distribution of aqueous U species varied greatly among exposure

223

waters. For example, in MOD water at 100 nM total U, speciation calculations show that Ca-U

224

ternary complexes, namely CaUO2(CO3)3-2 and Ca2UO2(CO3)3, dominate when pH exceeds 7.0

225

(Figure S3, Table 2). At pH 6.5, the abundance of Ca-U-ternary complexes decreased to reach

226

proportions similar to that for the binary U carbonate complexes, UO2CO3 and UO2(CO3)2-2.

227

Below pH 6.0, UO2CO3 dominates the fraction of total dissolved U. Across all pHs, the free

228

uranyl ion is present in negligible proportion. Adding DOM (14 mg C L-1) to MOD water at pH

229

7.5 decreased the fraction of U complexed as ternary Ca-U-CO3 species from 93 to 28% (Table

230

2), as DOM complexes 67% of the U. The proportion of binary U carbonate species also

231

decreased because of competition by the DOM for dissolved U.

232

Uranium efflux (experiment 5). 60% of the accumulated U was retained after 6 days of

233

depuration (Figure S4). The individual variability in the amount of U retained was large,

234

especially during the first 4 days of depuration. Loss of U after day 6 was not detectable.

235

Nonlinear regression fit of the mean values to equation 1 yielded k1 of 0.17 d-1, C1 of 40% and C2

236

of 60 %. The unidirectional efflux rate constant for the slow exchanging compartment (k2) could

237

not be inferred (was practically zero).

238

Discussion

239

Which aqueous U species best predicts bioavailability? The bioavailability of dissolved U

240

for L. stagnalis was not influenced by H+ competition for binding sites40 (Equation S5,

241

Supporting Information). Rather it was inversely related to the Ca and Mg concentrations of the

242

exposure waters (Figure 2). For instance, the kuw in very soft water is nearly 6-times greater than

243

in the harder synthetic freshwater. To explain the reduced bioavailability of U in harder waters,


11


Page 12 of 32

244

we first explored the possibility that Ca2+ and/or Mg2+, both hardness cations present in high

245

concentrations in hard freshwaters, competes with UO2+2 for biological uptake sites, thereby

246

reducing U bioaccumulation. To account for this competition, we modified the equations of

247

Croteau et al.40 and describe U bioaccumulation in L. stagnalis according to the free-ion activity

248

model when competition by Ca2+ or Mg2+ reduces U uptake (Equation S6, Supporting

249

Information). Taking into account competition between either Ca+2 or Mg2+ and UO2+2 for

250

binding sites did not, however, improve the predictive power of the relationship between the free

251

uranyl ion concentrations and the U concentrations in the snail.

252

Alternatively, the reduced bioavailability of U in hard freshwaters could be explained by a

253

lower abundance of “key” U species in harder than in softer waters. As shown in Table 2 and

254

Figure 3, decreasing the concentrations of hardness cations or decreasing the pH of MOD water

255

increased the proportion of binary carbonate species, with a concomitant decrease in the

256

proportion of ternary complexes. At pH 7.5, the fraction of all binary U carbonate species

257

increased from 5% for the harder synthetic freshwater (Kanab Creek) to 83% in the very soft

258

water (Table 2). The relative abundance of the carbonate species, which includes UO2(CO3)2-2

259

and to a much lesser extent (UO2)3(CO3)6-6 (more precisely, 8 to 21 orders of magnitude lower:

260

Table 3), is greater in very soft water where U is more bioavailable than in “hard” freshwaters

261

(Table 2). Ternary U complexes such as MgUO2(CO3)3-2 and CaUO2(CO3)3-2 dominate in harder

262

freshwaters where U is less bioavailable. These results thus suggest that the reduced

263

bioavailability of U in hard freshwaters may be explained by the low abundance of binary U-

264

carbonate species. The pH dependent relationship between kuw and the fractions of U species

265

further supported the importance of U-carbonate complex as a predictor of U bioavailability. As

266

shown in Figure 3, kuw is lower at pH 6.0 than at pH 6.5, and thus similar to the pH trend of the


12

Page 13 of 32


267

fraction of UO2(CO3)2-2. Ternary U complexes are most abundant at pH > 7, which coincides

268

with the lowest U bioavailability, as reflected by the low values of kuw, suggesting that the

269

ternary species are less bioavailable.

270

Across all water types, the uranium influx rate was highly dependent on the total U

271

concentration (R2 = 0.85 F(1,32)= 163; P < 0.001), as well as the concentration of most of the

272

computed U species. Given the redundancy among a large number of U species, using multiple

273

regression to identify bioavailable U species was problematic. Therefore, the identity of the best

274

predictor(s) of U bioavailability was revealed based on the regression coefficients of

275

relationships between U uptake rates into L. stagnalis and each of 23 U species present in the

276

experimental freshwaters (data for experiments 1-3). Two binary U-carbonate complexes,

277

namely UO2(CO3)2-2 and (UO2)3(CO3)6-6, each predict more than 90% of the variation in U

278

uptake rates (Figure 5A, Table 3). Based on the relative concentrations of these U-carbonate

279

complexes and the slightly greater R2 for UO2(CO3)2-2 than for (UO2)3(CO3)6-6 (Table 3), the

280

former U species appears as the best predictor of U bioavailability for L. stagnalis. These two U-

281

carbonate complexes predicted equally well U uptake rates when DOM was added (open circles,

282

Figure 5A), providing further evidence of the predictive capability of these U species. In

283

contrast, the free uranyl ion concentration ([UO2+2]) poorly predicted U uptake rates into L.

284

stagnalis (36%, Table 3, Figure 5B), challenging the FIAM postulate.

285

Can physiology explain in part the high U uptake rates in very soft freshwater? We tested

286

the hypothesis that high Ca influx rates explain in part the high U uptake rates observed in L.

287

stagnalis at low water hardness (Figure 2). Calcium requirement is extremely high in shell

288

forming freshwater mollusks41. As a result, Ca uptake rates should be enhanced in very soft

289

waters to compensate the low waterborne Ca concentrations. This hypothesis assumes that U


13


Page 14 of 32

290

uptake occurs through a membrane carrier dedicated to Ca7. However, there was no relationship

291

between either U uptake rates (Figure S5), or kuw (Figure 4a)and Ca uptake rates because Ca

292

uptake rates rapidly saturate with increasing Ca concentrations. But U bioavailability (inferred

293

from kuw) co-varied with the relative distribution of UO2(CO3)2-2 in the experimental medias,

294

including the highest U bioavailability observed at ~80 µM Ca (Figure 4b). Thus speciation, and

295

not physiology, explains the high U uptake rates in very soft freshwater. Binary U-carbonate

296

species, namely UO2(CO3)2-2 appear to predict aqueous U bioavailability in L. stagnalis.

297

Furthermore, these results suggest that U does not use a Ca membrane transporter exclusively,

298

and that other types of transporters are being used. Anionic species of U (Table 3) might, for

299

example, cross membranes via anionic channels as analogues for phosphate or bicarbonate, as

300

anion channels are often large and nonselective10. Uranium fluxes through anion transporters are

301

however expected to be low, considering the low U concentrations and because anion

302

transporters normally have relatively low affinities. Comparing our results to Cl- influx reported

303

for the species41 reveals that U uptake rates are 2 orders of magnitude slower than those for Cl-.

304

The kuw for U in MOD water is furthermore at the low end of those reported for other metals in

305

the species, which might reflect the low affinity of the U transporters (Table S7).

306

Process-based knowledge is needed to understand U bioavailability, to delineate exposure

307

pathways and to forecast potential environmental risks of U contamination. Our study underlined

308

the importance of aqueous speciation for U bioavailability. Given the link between

309

bioaccumulation and toxicity, exposures to aqueous U at low pH, low hardness and without

310

DOM, are more likely to elicit adverse effects more readily than equivalent U exposures at

311

elevated pH, elevated hardness or in the presence of DOM. Waterborne U uptake potentially

312

represents an important exposure pathway resulting from dissolution and/or desorption of U from


14

Page 15 of 32


313

U(VI)-bearing particles42,43. However dissolution of U from soluble U-bearing minerals, for

314

example during precipitation events, could change the partitioning of U between water and

315

sediment (or biofilm)34, affecting the U exposure pathways. Consumption of periphyton can be

316

an important route of metal exposure to benthic invertebrates44, however dietary U uptake

317

remains scarcely studied39. Future avenues of research include 1) assessment of the dietary

318

bioavailability of different forms of particulate U to allow delineation of U uptake routes; 2)

319

localization and characterization of the forms of U retained in tissues to get insights into the

320

mechanisms involved in U detoxification45; and 3) measurement of U concentrations in resident

321

organisms across a range of geochemical conditions such as a comparative study in springs

322

outflows in the Grand Canyon region to identify exposure pathways and assess the potential risks

323

associated with U exposure to native fauna.

324

ASSOCIATED CONTENT

325

Supporting Information. Snails’ size, dissolved U concentrations, pCO2 and pHs for the

326

experiments 1-3, chemical properties of DOM isolate, dissolved U concentrations, pHs and Ca

327

concentrations for the experiment 4, details for modeling U elimination and determining U-DOM

328

conditional binding constants, metal binding characteristics and rate constants, and graphical

329

illustrations of relationships between snail’s dry weight and time, U uptake rate and exposures, U

330

speciation versus pH, proportional loss of U and time, as well as equations for U

331

bioaccumulation according to the free-ion activity model for competition by Ca+2 and by H+.

332

This material is available free of charge via the Internet at http://pubs.acs.org.

333

AUTHOR INFORMATION

334

Corresponding Author


15


Page 16 of 32

335

*email: [email protected]

336

Author Contributions

337

The manuscript was written through contributions of all authors. All authors have given approval

338

to the final version of the manuscript.

339

ACKNOWLEDGMENT

340

This study was supported by the National Research Program including funds for a Topical

341

Research Team and the Toxics Substances Hydrology Program of the U.S. Geological Survey.

342

Kathy Akstin and David Barasch performed some of the ICP-MS analyses. Comments by Katie

343

Walton-Day, Jo Ellen Hinck and anonymous reviewers improved the manuscript. Any use of

344

trade, product, or firm names is for descriptive purposes only, and does not imply endorsement

345

by the U.S. Government.

346

REFERENCES

347

1. U.S. Bureau of Land Management. 2011. Northern Arizona proposed withdrawal final

348

environmental

impact

statement.

BLM/AZ/PL-11/002.

Volumes

I

and

349

http://www.blm.gov/az/st/en/info/nepa/environmental_library/eis/naz-withdraw.html.

II.

350

2. Van Dam, R.A.; Trenfield, M.A.; Markich, S.J.; Harford, A.J.; Humphrey, C.L.; Hogan,

351

A.C.; Stauber, J.L. Reanalysis of uranium toxicity data for selected freshwater organisms and

352

the influence of dissolved organic carbon. Environ. Toxicol. Chem. 2012, 31(11), 2606-2614.

353

3. Mathews, T.; Beaugelin-Seiller, K.; Garnier-Laplace, J.; Gilbin, R.; Adam, C.; Della-

354

Vedova, C. A probabilistic assessment of the chemical and radiological risks of chronic


16

Page 17 of 32


355

exposure to uranium in freshwater ecosystems. Environ. Sci. Technol. 2009, 43(17), 6684-

356

6690.

357 358

4. Swanson, S.M. Food-chain transfer of U-series radionuclides in a northern Saskatchewan aquatic system. Health physics, 1985, 49(5), 747-770.

359

5. Simon, O.; Garnier-Laplace, J. Kinetic analysis of uranium accumulation in the bivalve

360

Corbicula fluminea: effect of pH and direct exposure levels. Aquat. Toxicol. 2004, 68(2), 95-

361

108.

362 363

6. Suzuki, Y.; Banfield, J.F. Geomicrobiology of uranium. Rev. Miner. Geochem. 1999, 38(1), 393-432.

364

7. Köster, W. Transport of solutes across biological membranes: prokaryotes. In:

365

Physicochemical kinetics and transport at biointerfaces; Van Leeuwen, H.P. Köster, W.

366

Eds.; John Wiley & Sons, 2004, pp 271-335.

367

8. El Hage Chahine, J.-M.; Hémadi, M.; Ha-Duong, N.-T. Uptake and release of metal ions by

368

transferrin and interaction with receptor 1. Biochim. Biophys. Acta 2012, 1820(3): 334-347.

369 370

9. Muscatello, J.R.; Liber, K. Uranium uptake and depuration in the aquatic invertebrate Chironomus tentans. Environ. Poll. 2010, 158(5), 1696-1701.

371

10. Simkiss, K.; Taylor, M.G. Transport of metals across membranes. In: Metal speciation and

372

bioavailability in aquatic systems, Tessier A., Turner, D. R., Eds.; John Wiley & Sons, New

373

York, 1995, pp 1-44.


17


Page 18 of 32

374

11. Charles, A.L.; Markich, S.J.; Stauber, J.L.; De Filippis, L.F. The effect of water hardness on

375

the toxicity of uranium to a tropical freshwater alga (Chlorella sp.). Aquat. Toxicol. 2002,

376

60(1), 61-73.

377

12. Fortin, C.; Denison, F.H.; Garnier‐Laplace, J. Metal‐phytoplankton interactions: Modeling

378

the effect of competing ions (H+, Ca2+, and Mg2+) on uranium uptake. Environ. Toxicol.

379

Chem. 2007, 26(2), 242-248.

380

13. Trenfield, M.A.; Ng, J.C.; Noller, B.N.; Markich, S.J.; Dam, R.A.V. Dissolved organic

381

carbon reduces uranium bioavailability and toxicity. 2. Uranium [VI] speciation and toxicity

382

to three tropical freshwater organisms. Environ. Sci. Technol. 2011, 45(7), 3082-3089.

383

14. Campbell, P. G. C. Interactions between trace metals and aquatic organisms: A critique of the

384

free-ion activity model. In: Metal speciation and bioavailability in aquatic systems, Tessier

385

A., Turner, D. R., Eds.; John Wiley & Sons, New York, 1995, pp. 45-102.

386

15. Fortin, C.; Dutel, L.; Garnier‐Laplace, J. Uranium complexation and uptake by a green alga

387

in relation to chemical speciation: the importance of the free uranyl ion. Environ. Toxicol.

388

Chem. 2004, 23(4), 974-981.

389

16. Markich, S.J.; Brown, P.L.; Jeffree, R.A. The use of geochemical speciation modelling to

390

predict the impact of uranium to freshwater biota. Radiochim. Acta, 1996, 74(s1), 321-326.

391

17. Dong, W.; Brooks, S.C. Determination of the formation constants of ternary complexes of

392

uranyl and carbonate with alkaline earth metals (Mg2+, Ca2+, Sr2+, and Ba2+) using anion

393

exchange method. Environ. Sci. Technol. 2006, 40(15), 4689-4695.


18

Page 19 of 32


394

18. Lietsch, C.; Hoth, N.; Kassahun, A. Thermodynamic characterization of dissolved uranium

395

species in flooded uranium mines and tailing management facilities. In, An Interdisciplinary

396

Response to Mine Water Challenges, Sui, W., Sun, Y., Wang, C. Eds.; Proceedings of the

397

12th International Mine Water Association Congress (IMWA12), 2014, 137-141.

398

19. Taylor, H. E.; Peart, D. B.; Antweiler, R. C.; Brinton, T. I.; Campbell, W. L.; Garbarino, J.

399

R.; Roth, D. A.; Hart, R. J.; Averett, R. C. Data from synoptic water-quality studies on the

400

Colorado river in the grand canyon, Arizona, November 1990 and June 1991. USGS Open-

401

File Report 96-614, 1996, 175p

402

20. Prat, O.; Vercouter, T.; Ansoborlo, E.; Fichet, P.; Perret, P.; Kurttio, P.; Salonen, L. Uranium

403

speciation in drinking water from drilled wells in southern Finland and its potential links to

404

health effects. Environ. Sci. Technol. 2009, 43, 3941–3946.

405

21. Hummel, W., Mompean, F. J., Illemassène, M., Perrone, J. Chemical thermodynamics of

406

compounds and complexes of U, Np, Pu, Am, Tc, Se, Ni and Zr with selected organic

407

ligands, Vol 9, Elsevier Science, 2005, 1087 p

408 409

22. Croteau, MN.; Luoma, SN. 2009. Predicting dietborne metal toxicity from metal influxes. Environ. Sci. Technol. 2009, 43, 4915-4921.

410

23. Cain, DJ.; Croteau, MN.; Fuller, CC.; Ringwood, AH. Dietary uptake of Cu sorbed to

411

hydrous iron oxide is linked to cellular toxicity and feeding inhibition in a benthic grazer.

412

Environ. Sci. Technol. 2016, 50, 1552-1560.


19


Page 20 of 32

413

24. Alpine, A.E. Ed., Hydrological, geological, and biological site characterization of breccia

414

pipe uranium deposits in northern Arizona. U.S. Geological Survey Scientific Investigations

415

Report 2010–5025, 2010, 353 p.

416

25. U.S. Environmental Protection Agency. 2002. Methods for measuring the acute toxicity of

417

effluents and receiving waters to freshwater and marine organisms. EPA-821-R-02.

418

Washington DC

419

26. Stoiber, T.; Croteau, M.N.; Römer, I.; Tejamaya, M.; Lead, J.R.; Luoma, S.N. Influence of

420

hardness on the bioavailability of silver to a freshwater snail after waterborne exposure to

421

silver nitrate and silver nanoparticles. Nanotoxicology, 2015, 9(7), 918-927.

422

27. De With, ND.; Kamerik, G.; Slootstra, JW., Bergema, WF.; van Der Schors, RC. The effects

423

of external acidification on the ionic composition of the haemolymph, acid-base balance and

424

sodium metabolisms in the pulmonate freshwater snail, Lymnaea stagnalis. Comp. Biochem.

425

Physiol. 1989, 93A(4): 833-837.

426

28. Aiken, G. R.; McKnight, D. M.; Thorn, K. A.; Thurman, E. M. Isolation of hydrophilic

427

organic acids from water using nonionic macroporous resins. Org. Geochem. 1992, 18(4),

428

567–573.

429

29. Luoma, S.N.; Fisher, N.S. Uncertainties in assessing contaminant exposure from sediments.

430

In: Ecological risk assessments of contaminated sediments; Ingersoll, C.G., Dillon, T.,

431

Biddinger, G., R., Eds.; SETAC Special Publication Series: Pensacola, Fl, 1997.

432

30. Parkhurst, D.L.; Appelo, C.A.J. Description of input and examples for PHREEQC version

433

3—A computer program for speciation, batch-reaction, one-dimensional transport, and


20

Page 21 of 32


434

inverse geochemical calculations: U.S. Geological Survey Techniques and Methods, book 6,

435

chap. A43, 2013, 497 p.

436

31. Guillaumont, R.; Fanghängel, T.; Fuger, J.; Grenthe, I.; Neck, V.; Palmer, D. A.; Rand, M.

437

H. Update on the chemical thermodynamics of uranium, neptunium, plutonium, americium

438

and technetium, Vol. 5, Elsevier: Amsterdam, 2003, 919 p.

439

32. Dong, W.M., Brooks, S.C., 2008. Formation of aqueous MgUO2(CO3)32- complex and

440

uranium anion exchange mechanism onto an exchange resin. Environ. Sci. Technol. 2008, 42,

441

1979–1983.

442

33. Vercouter, T.; Reiller, P.E.; Ansoborlo, E.; Février, L.; Gilbin, R.; Lomenech, C.; Philippini,

443

V. A modelling exercise on the importance of ternary alkaline earth carbonate species of

444

uranium(VI) in the inorganic speciation of natural waters. Appl. Geochem. 2015, 55, 192–

445

198.

446

34. Stewart, B.D.; Mayes, M.A.; Fendorf, S. Impact of uranyl− calcium− carbonato complexes

447

on uranium (VI) adsorption to synthetic and natural sediments. Environ. Sci. Technol. 2010,

448

44(3), 928-934.

449

35. Nair S.; Karimzadeh L.; Merkel B.J. Surface complexation modeling of Uranium(VI)

450

sorption on quartz in the presence and absence of alkaline earth metals. Environ. Earth Sci

451

2013, 71, 1737-1745.

452

36. Bradbury, M.H., Baeyens, B., 2011. Predictive sorption modelling of Ni(II), Co(II), Eu(IIII),

453

Th(IV) and U(VI) on MX-80 bentonite and Opalinus Clay: a ‘‘bottom-up’’ approach. Appl.

454

Clay Sci. 2011, 52, 27–33.


21


455

Page 22 of 32

37. Fox, P.M.; Davis, J.A.; Zachara, J.M., 2006. The effect of calcium on aqueous uranium(VI)

456

speciation and adsorption to ferrihydrite and quartz. Geochim. Cosmochim. Acta 2006, 70,

457

1379–1387.

458

38. Van Loon, L. R.; Granacher, S.; Harduf, H. Equilibrium dialysis—ligand exchange: a novel

459

method for determining conditional stability constants of radionuclide—humic acid

460

complexes. Anal. Chim. Acta 1992, 268, 235–246.

461

39. Goulet, RR.; Fortin, C.; Spry, DJ. Uranium In: Homeostasis and toxicology of non-essential

462

metals; C. Wood, A. Farrell & C. Brauner [Eds]. 2012, 31B, Academic Press, pp391-428.

463

40. Croteau, M.N.; Hare, L.; Tessier, A. Refining and testing a trace metal biomonitor

464 465

(Chaoborus) in highly acidic lakes. Environ. Sci. Technol. 1998, 32(9), 1348-1353. 41. Grosell, M.; Brix, K.V. High net calcium uptake explains the hypersensitivity of the

466

freshwater pulmonate snail, Lymnaea stagnalis, to chronic lead exposure. Aquat. Toxicol.

467

2009, 91(4), 302-311.

468

42. Delemos, J.L.; Bostick, B.C.; Quicksall, A.N.; Landis, J.D.; George, C.C.; Slagowski, N.L.;

469

Rock, T.; Brugge, D.; Lewis, J.; Durant, J.L. Rapid dissolution of soluble uranyl phases in

470

arid, mine-impacted catchments near Church Rock, NM. Environ. Sci. Technol. 2008,

471

42(11), 3951-3957.

472

43. Blake, J.M.; Avasarala, S.; Artyushkova, K.; Ali, A.M.S.; Brearley, A.J.; Shuey, C.;

473

Robinson, W.P.; Nez, C.; Bill, S.; Lewis, J.; Hirani, C., Lezama Pacheco, J.S.; Cerrato, J.M.

474

Elevated Concentrations of U and Co-occurring Metals in Abandoned Mine Wastes in a

475

Northeastern Arizona Native American Community. Environ. Sci. Technol. 2015, 49(14),

476

8506-8514.


22

Page 23 of 32

477


44. Cain, D.J.; Croteau, M.-N.; Luoma, S. N. Bioaccumulation dynamics and exposure routes of

478

Cd and Cu among species of aquatic mayflies. Environ. Toxicol. Chem. 2011, 30, 2532-2541.

479

45. Mason, A.K.; Jenkins, K.D. Metal detoxification in aquatic organisms. In: Metal speciation

480

and bioavailability in aquatic systems, Tessier A., Turner, D. R., Eds.; John Wiley & Sons,

481

New York, 1995, pp 479-606.

482


23


Page 24 of 32

U uptake rate into L. stagnalis -1 -1 (nmole g d )

1000 VS SO MOD KC 100

10

1 controls

0.1 0.01

0.1

1

10

100

1000

10000

Dissolved U concentrations (nM)

483 484

Figure 1. Uranium uptake rate in L. stagnalis (soft tissues ± SD) exposed to aqueous U in

485

synthetic waters of different composition for 24 h (Experiment set 1). Each symbol represents

486

uptake rates for 10 individuals and 3 water samples. Symbols correspond to: very soft water

487

(VS), soft water (SO), moderately hard water (MOD), and Kanab Creek ASW (KC).


24

Page 25 of 32


1.8 VS

-1

U kuw

-1

U kuw (±95% CI, l g d )

1.5

1.2

(±95% CI, l g-1 d-1 )

1.8 1.5 1.2 0.9 0.6 0.3 0.0 0

0.9

1000

2000

3000

4000

Dis solve d M g conc. (µM)

SO

0.6

MOD

0.3

KC

0.0 0

400

800

1200

1600

Dissolved Ca concentrations (µM)

488 489

Figure 2. Uranium uptake rate constant (kuw) as a function of Ca concentrations in the exposure

490

water (Experiment set 1); VS = very soft, SO = soft. MOD = moderately hard, KC = Kanab

491

Creek ASW. Also shown in insert in the relationship between kuw and Mg concentrations in the

492

exposure water.


25


1.2

1.0 w/o DOM with DOM

U-CO3 1.0

0.8 Ca-U-CO3

0.8

0.6 0.6 0.4 0.4

0.2

UO2(CO3)2-2

Fraction of U species

U uptake rate constant (kuw, l/g/d ± SE)

Page 26 of 32

0.2

0.0

0.0 5.5

6.0

6.5

7.0

7.5

8.0

pH

493 494 495

Figure 3. Uranium uptake rate constant (± SE) as a function of pH in MOD water with (star) and

496

without (solid circles) DOM present (Experiment set 2 and experiment 3). Error bars smaller

497

than symbol for Experiment 3. Also shown is the computed fraction of total U as UO2(CO3)2-, U-

498

CO3 (the sum of U in all binary uranyl carbonate species (UO2(CO3)2-2, UO2CO3, UO2(CO3)3-4,

499

(UO2)2CO3(OH)3- and (UO2)3(CO3)6-6,) and Ca-U-CO3 (the sum of CaUO2(CO3)3-2 and

500

Ca2UO2(CO3)3 species.


26

Page 27 of 32


250

200

150

100

50

44

Ca influx in L. stagnalis (µmol/g/d)

300

0 0

100

200

300 0.4

3

0.3

2

0.2

1

0.1

0

fraction UO2(CO3)2

kuw U (± 95CI l/g/d)

-2

4

0.0 0

100

200

300

44

waterborne [ Ca] (µM)

501 44

502

Figure 4. A)

Calcium uptake rates in L. stagnalis soft tissues as a function of waterborne Ca

503

concentration (Experiment 4).

504

concentrations determined using the relative abundance of

505

inferred from 44Ca. The solid line represents the nonlinear regression fit of the mean values to a

506

one site ligand saturation model; B) Provisional rate constant of uptake for U as a function of the

507

waterborne Ca concentration. The error bars represent ±SD relative to the U concentration at

508

time 0. The dashed line represents the relative abundance of UO2(CO3)2-2 in the experimental

509

waters.

44

Calcium uptake rates were calculated using the total 44

44

Ca

Ca and the Ca concentrations


27


U uptake rates into L. stagnalis -1 -1 (nmole g d )

1000

Page 28 of 32

w/o DOM w/ DOM

100

A 10

1

0.1 0.01

0.1

1

10

100

1000

UO2(CO3)2-2 (nM)

U uptake rates into L. stagnalis -1 -1 (nmole g d )

1000

100

B 10

1

0.1 10-8

10-7

10-6

10-5

10-4

10-3

10-2

10-1

100

[UO2+2] (nM)

510 511 512

Figure 5. Uranium uptake rates (± SD) in L. stagnalis soft tissues as a function of the

513

concentrations of A) UO2(CO3)2-2 R2 = 0.92 B) UO2+2 R2 = 0.36; slopes and y-intercepts are

514

presented in Table 3.


28

Page 29 of 32


515

Table 1. Chemical composition (mM) of the

516

experimental media for each experiment; pH values

517

are average final pH; see Table S2 for range of U

518

concentrations; Very Softa

Softa

MODa Kanab Creekb

Na

0.14

0.57

1.14

1.48

Ca

0.044

0.17

0.35

1.45

Mg

0.062

0.25

0.50

2.96

K

0.007

0.027

0.054

0.16

SO4-2

0.11

0.42

0.85

6.03

Cl

0.007

0.027

0.054

0.44

Alkalinity 0.13

0.58

1.18

0.62

pH

7.4

6-8c

7.3

6.9

519

a

520 521 522 523 524 525

b

The Kanab ASW is based on major ion chemistry of samples from the mouth of Kanab Creek19. Because stream water was supersaturated with respect to calcite at atmospheric pCO2, dissolved Ca and carbonate alkalinity concentrations in the artificial stream water recipe were chosen to attain conditions below calcite saturation at the initial experimental pH. Na and Cl were increased to maintain the ionic strength equal to the stream water; Mg concentration was kept at the average measured in the stream.

526

c

Reference 25

pH controlled by imposing elevated pCO2

527


29


528

Page 30 of 32

Table 2. Species fraction of total U (25 nM) at pH 7.5. Species fraction Kanab of total U Creek

MOD

Soft

Very Soft

f(UO2+2)

9.3E-07

3.6E-07 2.1E-07

2.3E-05

1.2E-04

f(U-CO3)a

0.05

0.07

0.02

0.34

0.83

f(Ca,Mg)-U-CO3b

0.95

0.93

0.28

0.64

0.06

f(U-DOM)c 529

a

530

b

531

c

MOD + DOM

0.67

fraction of all U binary carbonate species; fraction of U ternary Ca and Mg carbonate species;

fraction of U complexed by DOM

532


30

Page 31 of 32


533

Table 3. Squared multiple R (R2), slopes and y-intercepts (± 95% CI) for the linear regressions

534

between U uptake rates into L. stagnalis (nmole g-1 d-1) and the concentrations of each U species

535

(M), including total U concentration (M), for experiments 1-3; n = 39, data log10-transformed; P

536

< 0.001, except for species #2 where P < 0.05. Also given is the concentration range for each U

537

form; * n = 35 Species #

U species

R2

slope

y-intercept

0 1 2 3 4 5 6 7* 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22

Total U UO2+2 CaUO2(CO3)3-2 Ca2UO2(CO3)3 MgUO2(CO3)3-2 UO2(CO3)2-2 UO2CO3 UO2(CO3)3-4 (UO2)2CO3(OH)3(UO2)3(CO3)6-6 UO2OH+ UO2(OH)2 UO2(OH)3(UO2)2(OH)2+2 UO2(OH)4-2 (UO2)3(OH)5+ (UO2)2(OH)+3 (UO2)3(OH)4+2 (UO2)3(OH)7(UO2)4(OH)7+ UO2Cl+ UO2Cl2 UO2SO4 UO2(SO4)2-2

0.85 0.36 0.53 0.11 0.54 0.92 0.69 0.65 0.71 0.91 0.64 0.80 0.77 0.65 0.58 0.74 0.51 0.71 0.77 0.74 0.41 0.35 0.44 0.40

0.702 ± 0.100 0.302 ± 0.135 0.502 ± 0.157 0.173 ± 0.163 0.481 ± 0.148 0.778 ± 0.077 0.548 ± 0.122 0.552 ± 0.136 0.287 ± 0.064 0.258 ± 0.028 0.466 ± 0.117 0.542 ± 0.091 0.511 ± 0.092 0.229 ± 0.028 0.331 ± 0.094 0.182 ± 0.036 0.163 ± 0.053 0.170 ± 0.037 0.136 ± 0.025 0.130 ± 0.026 0.324 ± 0.128 0.279 ± 0.128 0.352 ± 0.132 0.336 ± 0.138

6.38 ± 0.76 4.96 ± 1.75 5.22 ± 1.32 2.59 ± 1.45 5.46 ± 1.36 7.92 ± 0.68 6.16 ± 1.14 6.29 ± 1.29 4.42 ± 0.75 6.62 ± 0.60 6.18 ± 1.29 6.86 ± 0.98 6.99 ± 1.07 4.94 ± 0.96 6.48 ± 1.54 4.42 ± 0.68 4.47 ± 1.12 4.78 ± 0.81 4.23 ± 0.59 4.06 ± 0.61 6.60 ± 2.20 7.42 ± 2.91 5.66 ± 1.72 6.18 ± 2.11

Conc. range (M) 10-12.0 – 10-5.28 10-17.3 – 10-9.68 10-14.1 – 10-5.54 10-15.1 – 10-5.63 10-15.0 – 10-6.43 10-12.9 – 10-6.40 10-13.5 – 10-6.82 10-14.9 – 10-6.84 10-19.2 – 10-5.67 10-34.1 – 10-14.5 10-15.1 – 10-7.74 10-14.5 – 10-7.43 10-15.2 – 10-8.3 10-25.2 – 10-10.6 10-20.3 – 10-9.29 10-30.2 – 10-8.61 10-29.7 – 10-7.18 10-33.9 – 10-12.1 10-31.9 – 10-10.8 10-38.9 – 10-10.2 10-22.5 – 10-14.1 10-29.1 – 10-19.2 10-17.7 – 10-10.6 10-20.1 – 10-11.9

538


31


539

Page 32 of 32

Table of Contents Graphic

540


32

Biogeochemical Controls of Uranium ... - ACS Publications

Recommend Documents