Koninklijke Bibliotheek, National Library of the Netherlands
IP1554970138112
Speciation dynamics of oxyanion contaminants (As, Sb, Cr) in argillaceous suspensions during oxic-anoxic cycles
Markelova, Ekaterina
Couture, Raoul-Marie
Parsons, Christopher T.
Markelov, Igor
Madé, Benoit
Van Cappellen, Philippe
Charlet, Laurent
text
article
monographic
Applied geochemistry
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Elektronische Wetenschappelijke Tijdschriften
EWTIJ
10.1016/j.apgeochem.2017.12.012
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10.1016/j.apgeochem.2017.12.012
The Authors
Fig. 1
Time series of measured redox potential (EH) (line) and pH (dashed line) values (a–d), nitrate (NO3
−) (down-pointing triangle), nitrite (NO2
−) (up-pointing triangle), and ATP concentrations (squares connected by dashed line) (e–h). Modelled concentrations of NO3
− and NO2
− are indicated by solid lines. Addition of microbial inhibitors is indicated by stars (e), whereas shaded and white areas indicate anoxic (CO2 + N2) and oxic (O2 + CO2 + N2) periods, respectively.
Fig. 1
Fig. 2
Time series of modelled (line) and measured [As] (diamond), [Sb] (circle), and [Cr] (square) aqueous concentrations, (a–d); measured concentrations of dissolved inorganic carbon (DIC) (down-pointing triangle) and carbon in the form of acetate (up-pointing triangle) (e–h).
Fig. 2
Fig. 3
Time series of measured total aqueous concentrations (circle) and solid speciation (star) of As (a–d), Sb (e–h), and Cr (i–l). Reaction formulations and parameters used in the models to predict aqueous (solid line) and solid (dashed line) speciation are given in Table 2 and described in Appendix D.
Fig. 3
Fig. 4
Redox scale showing the theoretical succession of redox potentials (EH in mV) calculated for the initial experimental conditions (pH 6.9, activities of oxidized and reduced species; Table 4). Potentially toxic CrO4
2−, HAsO4
2−, and Sb(OH)6
− can be reduced simultaneously, likely via detoxification, with no dependency on EH (on the left) under aerobic and denitrifying conditions, or successively as terminal electron acceptors (TEAs) via respiration following the thermodynamic order (i.e., EH cascade) (on the right). Note that NO3
− is exclusively reduced via respiration in the presence of microorganisms.
Fig. 4
Fig. 5
Contribution of the various reaction pathways to the simultaneous sequestration of aqueous As(V), Sb(V), and Cr(VI) simulated by the model for the experimental end points. The results of Exp. III where the native microbial community was stimulated by the addition of organic carbon are similar to those of Exp. II within 5%, and are therefore not shown.
Fig. 5
Table 1
Design of batch experiments comprising the Tégulines sample (50 g L−1) dispersed in 1 L of synthetic pore water (Table B.1) and amended with a mixture of As(V), Sb(V), and Cr(VI).
Table 1
Factors tested
Exp. I
Exp. II
Exp. III
Exp. IV
Sterile mineralogy
+
+
+
+
Microbial activity of native community
+
+
+
Labile organic carbon (ethanol) addition
+
+
Microbial activity of enriched (native + inoculated) community
+
Table 2
Reaction network and kinetic formulations used in the biogeochemical models. Parameter values are given in Annex D (Table D.2).
Table 2
Description
Reaction
Kinetic or equilibrium formulation
R#
Electrode leakage
R
1
=
r
l
e
a
k
D
O
C
1
Fermentation of ethanol
C2H5OH + H2O → CH3COO− + 5 H+ + 4 ē
R
2
=
k
C
2
H
5
O
H
⋅
[
C
2
H
5
O
H
]
⋅
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
2
Fermentation of acetate
CH3COO− + 4 H2O →2 HCO3
− + 9 H+ + 8 ē
R
3
=
k
C
H
3
C
O
O
−
⋅
[
C
H
3
C
O
O
−
]
⋅
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
3
Aerobic respiration of ethanol
O2 + C2H5OH →CH3COO− + H2O + H+
R
4
=
k
e
t
h
O
2
⋅
[
C
2
H
5
O
H
]
⋅
(
[
O
2
]
[
O
2
]
+
K
m
O
2
)
4
Aerobic respiration of acetate
2 O2 + CH3COO− + H+ →2 CO2 + 2 H2O
R
5
=
k
a
c
O
2
⋅
[
C
H
3
C
O
O
−
]
⋅
(
[
O
2
]
[
O
2
]
+
K
m
O
2
)
5
Arsenic
Equilibrium adsorption of As(V)
HAsO4
2− + 2 ≡SOH ⇄≡SOHAsO4 + H2O
l
o
g
K
s
o
r
p
H
A
s
O
4
2
−
6
Kinetic sorption of As(V)
HAsO4
2− + 2 ≡SOH ⇄≡SOHAsO4 + H2O
R
f
7
=
k
f
H
A
s
O
4
2
−
⋅
[
H
A
s
O
4
2
−
]
7
R
b
8
=
k
b
H
A
s
O
4
2
−
⋅
≡
S
O
H
A
s
O
4
8
R
i
9
=
k
i
H
A
s
O
4
2
−
⋅
≡
S
O
H
A
s
O
4
9
Detoxification of aqueous As(V)
HAsO4
2− + 4 H+ + 2 ē →H3AsO3 + H2O
R
d
e
t
o
x
10
=
r
m
a
x
H
A
s
O
4
2
−
⋅
(
[
H
A
s
O
4
2
−
]
K
m
(
d
e
t
o
x
)
H
A
s
O
4
2
−
+
[
H
A
s
O
4
2
−
]
)
10
Description
Reaction
Kinetic or equilibrium formulation
R#
Respiration of aqueous As(V)
4 HAsO4
2− + C2H4O2 + 6 H+ → 4 H3AsO3 + 2 HCO3
−
R
r
e
s
p
11
=
k
r
e
s
p
H
A
s
O
4
2
−
⋅
[
C
H
3
C
O
O
−
]
⋅
(
[
H
A
s
O
4
2
−
]
K
m
(
r
e
s
p
)
H
A
s
O
4
2
−
+
[
H
A
s
O
4
2
−
]
)
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
11
Equilibrium adsorption of As(III)
H3AsO3 + 2≡SOH ⇄ ≡S2HAsO3 + 2 H2O
l
o
g
K
s
o
r
p
H
3
A
s
O
3
12
Kinetic sorption of As(III)
H3AsO3 + 2≡SOH ⇄≡S2HAsO3 + 2 H2O
R
f
13
=
k
f
H
3
A
s
O
3
⋅
[
H
3
A
s
O
3
]
R
b
14
=
k
b
H
3
A
s
O
3
⋅
≡
S
2
H
A
s
O
3
R
i
15
=
k
i
H
3
A
s
O
3
⋅
≡
S
2
H
A
s
O
3
131415
Oxidative biomethylation of As(III)
CH4 + H3AsO3 →CH3AsVO(OH)2 + 2 H+ + 2 ē
R
m
e
t
h
y
l
16
=
r
m
a
x
H
3
A
s
O
3
⋅
(
[
H
3
A
s
O
3
]
K
m
H
3
A
s
O
3
+
[
H
3
A
s
O
3
]
)
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
16
Antimony
Equilibrium adsorption of Sb(V)
Sb(OH)6
− + 2 ≡SOH ⇄(≡SO)2Sb(OH)4
− + 2 H2O
l
o
g
K
s
o
r
p
S
b
(
O
H
)
6
−
17
Kinetic sorption of Sb(V)
Sb(OH)6
− + 2 ≡SOH ⇄(≡SO)2Sb(OH)4
− + 2 H2O
R
f
18
=
k
f
S
b
(
O
H
)
6
−
⋅
[
S
b
(
O
H
)
6
−
]
18
R
b
19
=
k
b
S
b
(
O
H
)
6
−
⋅
(
≡
S
O
)
2
S
b
(
O
H
)
4
−
19
R
i
20
=
k
i
S
b
(
O
H
)
6
−
⋅
(
≡
S
O
)
2
S
b
(
O
H
)
4
−
20
Detoxification of surface-bound Sb
(≡SO)2Sb(OH)4
− + 3 H+ + 2 ē →Sb(OH)3 + 3H2O
R
d
e
t
o
x
21
=
r
m
a
x
S
b
(
O
H
)
6
−
⋅
(
(
≡
S
O
)
2
S
b
(
O
H
)
4
−
K
m
(
d
e
t
o
x
)
S
b
(
O
H
)
6
−
+
(
≡
S
O
)
2
S
b
(
O
H
)
4
−
)
21
Description
Reaction
Kinetic or equilibrium formulation
R#
Respiration of aqueous As(V)
4 HAsO4
2− + C2H4O2 + 6 H+ →4 H3AsO3 + 2 HCO3
−
R
r
e
s
p
11
=
k
r
e
s
p
H
A
s
O
4
2
−
⋅
[
C
H
3
C
O
O
−
]
⋅
(
[
H
A
s
O
4
2
−
]
K
m
(
r
e
s
p
)
H
A
s
O
4
2
−
+
[
H
A
s
O
4
2
−
]
)
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
11
Precipitation of Sb(III)
Sb2O3 (c) + 3 H2O ⇄2 Sb(OH)3 (aq)
l
o
g
K
s
p
S
b
2
O
3
23
Chromium
Equilibrium adsorption of Cr(VI)
CrO4
2− + ≡SOH + H+ ⇄ ≡SOH2
+CrO4
2−
l
o
g
K
s
o
r
p
C
r
O
4
2
−
24
Kinetic sorption of Cr(VI)
CrO4
2− + ≡SOH + H+ ⇄≡SOH2
+CrO4
2−
R
f
25
=
k
f
C
r
O
4
2
−
⋅
[
C
r
O
4
2
−
]
R
b
26
=
k
b
C
r
O
4
2
−
⋅
≡
S
O
H
2
+
C
r
O
4
2
−
R
i
27
=
k
i
C
r
O
4
2
−
⋅
≡
S
O
H
2
+
C
r
O
4
2
−
252627
Abiotic reduction of surface-bound Cr(VI)
≡SOH2
+CrO4
2− + 5 H+ + 3 ē →Cr(OH)3(s) + H2O
R
a
b
28
=
k
a
b
C
r
O
4
2
−
⋅
≡
S
O
H
2
+
C
r
O
4
2
−
28
Detoxification of aqueous Cr(VI)
CrO4
2− + 5 H+ + 3 ē →Cr(OH)3(s) + H2O
R
d
e
t
o
x
29
=
r
m
a
x
C
r
O
4
2
−
⋅
(
[
C
r
O
4
2
−
]
K
m
(
d
e
t
o
x
)
C
r
O
4
2
−
+
[
C
r
O
4
2
−
]
)
29
Description
Reaction
Kinetic or equilibrium formulation
R#
Respiration of aqueous Cr(VI)
8 CrO4
2− + 3 CH3COO− + 13 H+ ++4 H2O → 8 Cr(OH)3 + 6 HCO3
−
R
r
e
s
p
30
=
k
r
e
s
p
C
r
O
4
2
−
⋅
[
C
H
3
C
O
O
−
]
⋅
(
[
C
r
O
4
2
−
]
K
m
(
r
e
s
p
)
C
r
O
4
2
−
+
[
C
r
O
4
2
−
]
)
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
30
Precipitation of Cr(III)
Cr(OH)3(s) + H+ ⇄Cr(OH)2
+ + H2O
l
o
g
K
s
p
C
r
(
O
H
)
3
31
Nitrogen
Respiration of N(V)
2 NO3
− + CH2O →2 NO2
− + CO2 + H2O
R
r
e
s
p
32
=
k
N
O
3
−
⋅
[
C
H
2
O
]
⋅
(
[
N
O
3
−
]
K
m
N
O
3
−
+
[
N
O
3
−
]
)
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
(
K
i
n
C
r
O
4
2
−
K
i
n
C
r
O
4
2
−
+
[
C
r
O
4
2
−
]
)
32
Respiration of N(III)
4 NO2
− + 3 CH2O + 4 H+ →2 N2 + 3 CO2 + 5 H2O
R
r
e
s
p
33
=
k
N
O
2
−
⋅
[
C
H
2
O
]
⋅
(
[
N
O
2
−
]
K
m
N
O
2
−
+
[
N
O
2
−
]
)
(
K
i
n
O
2
K
i
n
O
2
+
[
O
2
]
)
(
K
i
n
C
r
O
4
2
−
K
i
n
C
r
O
4
2
−
+
[
C
r
O
4
2
−
]
)
33
Comments:
•≡SOH: a generic sorption site, which can be protonated or non-protonated and may involve one, two or three surface oxygens.
•CH2O: dissolved organic carbon released by the argillaceous matrix.
•Detoxification: microbial reduction of As, Sb and Cr, which is not coupled to the oxidation of an external electron donor.
•Respiration: microbial reduction of As, Sb and Cr, which is coupled to the oxidation of ethanol or acetate.
Table 3
Forward (kf), backward (kb) and irreversible (ki) rate constants, plus the equilibrium sorption constant (log K) used in the multi-reaction models. Estimated distribution (or partition) coefficients (Kd) of As(V), As(III), Sb(V), and Cr(V) for sorption to the Tégulines matrix in Exp. I.
Table 3
Sorbate
Kd (L kg−1)
kf (h−1)
kb (h−1)
ki (h−1)
Log K
RMSE
As(V)
46.7
3.0 × 10−3
5.8 × 10−3
2.3 × 10−3
−0.54
1.55
As(III)
–
1.7 × 10−5
4.8 × 10−2
2.7 × 10−3
−0.38
–
Sb(V)
9.8
1.1 × 10−3
4.2 × 10−3
6.7 × 10−4
−1.07
3.01
Cr(VI)
11108
4.0 × 10−3
3.8 × 10−2
0.0
−1.03
6.54
Table 4
Thermodynamic sequence based on the redox potentials (EH) of half reduction reactions at the experimental pH (6.9), temperature (25 °C), and activities of reduced and oxidized species at the beginning of Exp. IV. The thermodynamic order of As and Sb reduction differs depending on pH (EH°′ and EH°) and products of Sb reduction. The observed experimental sequence of Sb reduction followed by As reduction, suggests Sb reduction to Sb2O3(s), rather than to Sb(OH)3. Calculations are based on the thermodynamic values given in Bard et al. (1985); Filella and May (2003); Nordstrom et al. (2014); Tratnyek and Macalady (2000); Zotov et al (2003).
Table 4
Half-reduction reactions
EH [mV]pH 6.9
EH°′ [mV]pH 7
EH° [mV]pH 0
O2 + 4 H+ + 4 ē → 2 H2O
815
816
1230
CrO4
2− + 5 H+ + 3 ē → Cr(OH)3 + H2O
524
579
1270
NO3
− + 2 H+ + 2 ē → NO2
− + H2O
316
426
840
Sb(OH)6
− + 3 H+ + 2 ē → Sb(OH)3 + 3 H2O
188
91
712
HAsO4
2− + 4 H+ + 2 ē → H3AsO3 + H2O
160
18
846
2 Sb(OH)6
− + 6 H+ + 4 ē → Sb2O3(s) + 9 H2O
130
218
839
CO2 + 12 H+ + 12 ē → C2H5OH (ethanol) + 3 H2O
−795
−356
90
Speciation dynamics of oxyanion contaminants (As, Sb, Cr) in argillaceous suspensions during oxic-anoxic cycles
Ekaterina
Markelova
a
b
∗
emarkelo@uwaterloo.ca
ekaterina.markelova@ujf-grenoble.fr
Raoul-Marie
Couture
b
c
rmc@niva.no
Christopher T.
Parsons
b
chris.parsons@uwaterloo.ca
Igor
Markelov
a
b
igor.markelov@uwaterloo.ca
Benoit
Madé
d
Benoit.Made@andra.fr
Philippe
Van Cappellen
b
pvc@uwaterloo.ca
Laurent
Charlet
a
b
laurent.charlet@ujf-grenoble.fr
a
University Grenoble Alpes, University Savoie Mont Blanc, CNRS, IRD, IFSTTAR, ISTerre, 38000 Grenoble, France
University Grenoble Alpes
University Savoie Mont Blanc
CNRS
IRD
IFSTTAR
ISTerre
Grenoble
38000
France
b
Ecohydrology Research Group, Department of Earth and Environmental Sciences and Water Institute, University of Waterloo, 200 University Ave. W, Waterloo, ON N2L 3G1, Canada
Ecohydrology Research Group
Department of Earth and Environmental Sciences and Water Institute
University of Waterloo
200 University Ave. W
Waterloo
ON N2L 3G1
Canada
c
Norwegian Institute for Water Research-NIVA, Gaustadalléen 21, 0349 Oslo, Norway
Norwegian Institute for Water Research-NIVA
Gaustadalléen 21
Oslo
0349
Norway
d
Andra, National Radioactive Waste Management Agency, R&D Division, Transfer Migration Group, 1/7 rue Jean Monnet, 92298 Chatenay Malabry Cedex, France
Andra
National Radioactive Waste Management Agency
R&D Division
Transfer Migration Group
Chatenay Malabry Cedex
1/7 rue Jean Monnet
92298
France
∗
Corresponding author. University Grenoble Alpes, University Savoie Mont Blanc, CNRS, IRD, IFSTTAR, ISTerre, 38000 Grenoble, France.
University Grenoble Alpes
University Savoie Mont Blanc
CNRS
IRD
IFSTTAR
ISTerre
Grenoble
38000
France
Abstract
Argillaceous geological formations are considered promising repositories for waste containing inorganic contaminants. However, the sequestration capacity of an argillaceous natural barrier may change as a result of dynamic environmental conditions, in particular changes in redox state. Here, we imposed redox cycles to argillaceous suspensions amended with a mixture of the inorganic contaminants As(V), Sb(V), Cr(VI), by alternating 7-day cycles of sparging with either oxic or anoxic gas mixtures. During the redox cycles, we assessed the relative importance of different contaminant sequestration mechanisms under sterile and non-sterile conditions. The non-sterile experiments were carried out 1) with the argillaceous material as is, that is, containing the native microbial population, 2) with the addition of ethanol as a source of labile organic carbon, and 3) with the addition of both ethanol and a soil microbial inoculum. In order to capture the observed solid-solution partitioning trends, a series of mixed thermodynamic-kinetic models representing the dynamic conditions in the experiments were developed using PHREEQC. Parameter values for processes such as adsorption, which occurred in all experiments, were kept constant across all the simulations. Additional formulations were introduced as needed to account for the microbial detoxification and respiration of the contaminants in the increasing complexity experiments. Abiotic adsorption of As and abiotic reductive precipitation of Cr dominated sequestration of these two contaminants under all experimental conditions, whereas the presence of microorganisms was essential to decrease the aqueous Sb concentration. Once reduced, As(III), Sb(III), and Cr(III) were not re-oxidized upon exposure to dissolved O2 in any of the experiments. Neither the mineralogy nor the native microbiota catalysed contaminant oxidation processes. Thus, kinetically, the reduction of As(V), Sb(V), and Cr(VI) was irreversible over the experimental duration (49 days). The capacity of the argillaceous matrix to reduce and sequester the aqueous contaminants, without subsequent remobilization to the aqueous phase during the oxic periods, supports the use of argillaceous barriers for geologic waste storage.
Graphical abstract
Image 1
Highlights
•
Aqueous As(V) is removed by adsorption, Cr(VI) by abiotic reductive precipitation.
•
Microorganisms remove aqueous Sb(V) via reductive precipitation.
•
Once reduced, As(III), Sb(III) and Cr(III) are not oxidized when exposed to O2.
•
Oxyanions are reduced in parallel via detoxification, sequentially via respiration.
•
Measured EH does not respond quantitatively to observed speciation of Cr, N, As, Sb.
Keywords
Argillaceous suspension
Nuclear waste repository
Redox oscillations
Arsenic
Antimony
Chromium
Sequestration
Reduction
Respiration
Detoxification
1
Introduction
Argillaceous formations are being investigated as potential geological hosts for radioactive waste disposal in eleven Nuclear Energy Agency member countries, which are collectively referred to as the Clay Club (Nuclear Energy Agency, 1958). In low-level radioactive waste (e.g., long-lived radium-bearing waste) potentially toxic elements, including As, Sb, and Cr, are often present in addition to radionuclides, and may be released into the argillaceous porous media if failure of engineered barriers occurs (Andra, 2009). Furthermore, these contaminants are of general geochemical interest as they are often released to near-surface environments due to industrial pollution adjacent to mining and smelting operations (Cornelis et al., 2008; Wilson et al., 2010), shooting ranges (Johnson et al., 2005), and chemical weathering of natural rocks and minerals (e.g., pyroxenite, serpentinite, arsenopyrite, scorodite, stibnite; Godgul and Sahu, 1995; Wilson et al., 2010).
The mobility, bioavailability, and chemical reactivity of redox-sensitive As, Sb, and Cr are influenced by their speciation, which evolves according to changes in redox conditions, pH, electrolyte composition, temperature, and microbial activity (e.g., Fawcett et al., 2015). At near-neutral pH, oxidized species are expected to be present as oxyanions (primarily HAsO4
2−, Sb(OH)6
−, and CrO4
2−) with low sorption affinity to typically negatively charged clay surfaces dominant in an argillaceous matrix. Therefore, similarly to radionuclides, which may be present in anionic forms (e.g., Br, I) (Tournassat et al., 2007; Wersin et al., 2011), the potentially toxic contaminants As, Sb and Cr are of concern for waste management due to their expected mobility in the absence of positively charged Fe- or Al-oxide minerals (Manning and Goldberg, 1997; Zhang and Selim, 2005).
Dynamic changes in redox conditions have significant effects on aqueous contaminant concentrations in saturated near-surface environments (Couture et al., 2015; Hindersmann and Mansfeldt, 2014; Parsons et al., 2013; Stewart et al., 2009; Thompson et al., 2006), suggesting that they also play a role in controlling contaminant mobility in subsurface argillaceous formations. Research on how redox phenomena control the long-term mobility of radionuclides in waste repositories has been performed by 32 European research institutes and universities within the ReCosy collaborative project (Duro et al., 2014). However, in contrast to radionuclides, the mobility of associated metals and metalloids has received much less attention in waste management research.
Most of the experimental and modelling investigations assessing the long-term performance of geological repositories assume that highly reducing conditions develop and are maintained after repository closure (e.g., Hedin and Olsson, 2016; Laaksoharju et al., 2008). The assumption of stable reducing conditions may not always hold since oxidizing conditions can last from a few years to a hundred years during the initial period of repository utilization (Bildstein et al., 2005; Duro et al., 2014; Nagra, 2002; Stroes-Gascoyne et al., 2002). The intrusion of oxygenated groundwater (Swift and Bonano, 2016) or glacial melt water (Duro et al., 2014; Laaksoharju et al., 2008), as well as water radiolysis (Rodwell et al., 2003; Song and Zhang, 2008) are further potential causes of redox perturbations over the life-time of the repository. Therefore, the examination of natural barrier materials in contact with potentially toxic contaminants under redox-oscillating conditions can help inform scenarios of repository safety.
Besides the potential role of the redox conditions, contaminant mobility within repositories can be altered by the presence of microorganisms, which are commonly found to depths of at least 1500 m (Pedersen, 2002). In fact, microbial activity plays an important role in shaping geochemical processes even in deep, low-nutrient environments, such as granitic bedrock (Laaksoharju et al., 2008), and thus should not be ignored in environmental assessments of future repositories. A number of safety assessments already incorporate microbial processes into the predictive modelling of repository conditions (e.g., Hallbeck and Pedersen, 2012; Laaksoharju et al., 2008; Small et al., 2008). Prediction of contaminant mobility influenced by microbial activity should also be implemented.
The high heterogeneity of waste materials complicates environmental assessments, in particular when multiple contaminants are present in a mixture of coexisting redox-sensitive elements. Contaminated soils and waste waters are often found to contain high concentrations of As, Sb, and Cr as well as redox-sensitive nutrients, such as nitrogen (N) and organic carbon (OC) (Casiot et al., 2007; Hindersmann and Mansfeldt, 2014; Lindsay et al., 2011). Heterogeneous contaminant-nutrient mixtures are also prevalent in waste repositories (Bertron et al., 2014). Given that the biogeochemical cycles of nutrients and contaminants are intricately linked together (Andrewes et al., 2000; Couture et al., 2015; Stucker et al., 2014), there is a need for investigating multi-contaminant mixtures more systematically.
In this study, we evaluate the mobility of oxidized oxyanion contaminants (HAsO4
2−, Sb(OH)6
−, and CrO4
2−) in argillaceous suspensions under a range of variable environmental conditions. We present the results of a series of batch experiments focusing on the effects of oxic vs. anoxic and biotic vs. abiotic conditions in bioreactors. The role of adding labile OC (ethanol) as an electron donor, and the effect of microbial diversity are also evaluated. We hypothesized that (1) reintroduction of O2 in the argillaceous suspension would increase contaminant mobility via oxidation of reduced compounds, (2) increased heterotrophic microbial activity and ethanol addition would decrease contaminant mobility by increasing reductive precipitation of Sb and Cr, and (3) the contaminants would be reduced sequentially following the order of decreasing redox potential of the corresponding reduction reactions.
In the experiments, the complexity of the reactive systems was progressively increased. Observed geochemical patterns were interpreted using mixed kinetic-thermodynamic models that incorporated an increasing number of reaction pathways. Parameter values obtained for processes in simpler systems were fixed and applied unchanged in models of more complex systems, hence allowing us to assess the transferability of the reaction parameters. The models provide diagnostic and prognostic tools for the quantitative simulation of biogeochemical processes driving the partitioning of As, Sb, and Cr in argillaceous systems.
2
Materials and methods
2.1
Argillaceous matrix, synthetic pore water, and contaminant mixture
A sample of Tégulines argillaceous material from the Gault formation in north-eastern France was collected at 22.5 m depth and preserved under anaerobic conditions prior to use. The initial mineral composition of the argillaceous matrix was characterised by powder X-ray diffraction (XRD) using a Bruker Axs D8 diffractometer. Prior to analysis, the sample was dried undisturbed under an O2-free atmosphere in an anaerobic chamber (Jacomex BS 531, EP20Ra-nm purification, 100% N2, <1 ppm O2) and ground by hand in an agate mortar to a particle size of approximately 1–3 μm. Diffractograms were obtained using a Cu-target X-ray tube analysing Kα1+2 lines with a solXE detector. The X-ray diffraction patterns were treated by Rietveld refinement using BGMN software (Bergmann et al., 1998). Elemental composition was determined by inductively coupled plasma optical emission spectrometry (ICP-OES) (Varian 720 ES) after total acid digestion (HNO3 + HF + H3BO3 + H2O2) (Cotten et al., 1995). Concentrations of reducible Fe and Mn, as well as those of reactive, Al were estimated by performing citrate-bicarbonate-dithionite (CBD) extractions (Jackson et al., 1986).
After grinding, the argillaceous sample was sieved through a 1 mm mesh, dispersed in synthetic pore water (50 g L−1) and equilibrated, with agitation, for 7 days under O2-free conditions prior to oxyanion additions. The synthetic pore water (Table B.1) mimicked the solution composition in contact with the host rock formation in the absence of O2 (Gaucher et al., 2009). It was enriched in NO3
− and NH4
+ in order to simulate the leachate of a bitumen matrix (Bertron et al., 2014). The oxyanion mixture was prepared by dissolving high purity analytical grade Na2HAsO4⋅7H2O (≥98%, SIGMA), KSb(OH)6 (≥99%, Fluka), and K2CrO4 (≥99%, SIGMA-ALDRICH). Individual stock solutions of As and Cr were combined and added to the pre-equilibrated experimental suspensions, while Sb was spiked separately to prevent precipitation of NaSb(OH)6 due to the high stock concentration of Na (Ilgen et al., 2014).
2.2
Experiments
Batch experiments were performed using a bioreactor system based on the design of Thompson et al. (2006) and modified by Parsons et al. (2013). Briefly, each reactor contained a suspension volume of 1 L with a headspace of about 300 ml and incubated in the dark. Electrodes, mechanical agitator, and sampling ports were installed via the reactor head-plate while a water jacket ensured constant temperature (25 ± 1 °C). Solid polymer open junction pH and redox potential (EH) electrodes (Mettler-Toledo Xerolyt Solid) were connected to field-effect transistor (FET) instrumental amplifiers with high input impedance. The EH and pH data were recorded every 2 min using an Agilent 34970A Data Logger. Measured EH was corrected for the reference electrode's standard voltage (+207 mV) relative to the standard hydrogen electrode (SHE) at 25 °C and referred to a Ag/AgCl (3 M KCl) reference electrode.
Experimental time zero corresponds to the time of injection of the oxyanion mixture to a pre-equilibrated suspension under anoxic conditions. Redox oscillations were induced by alternate sparging of anoxic (1% CO2 + 99% N2) and oxic (1% CO2 + 79% N2 + 20% O2) gas mixtures for periods of 7 days for a total experimental time of 7 weeks. Sampling was performed on the 1st, 2nd, 4th, and 7th day of each anoxic and oxic period using gas-tight sterile syringes. Suspensions were subsampled, immediately centrifuged, and filtered through cellulose hydrophilic membrane filters (0.2 μm pore size, Chromafil RC, ROTH) in the anaerobic chamber. A part of the supernatant was flash frozen with liquid nitrogen and stored at −80 °C prior to aqueous speciation analysis. The rest of the supernatant was acidified with ultrapure HNO3 to pH < 3 and stored at 4 °C for total aqueous concentration analyses.
Each batch experiment included the Tégulines argillaceous material and the synthetic pore water (Table B.1) amended with the oxyanion mixture. The targeted initial contaminant concentrations were 500 μM HAsO4
2−, 500 μM Sb(OH)6
−, and 500 μM CrO4
2− (Table B.2). In total, four increasing complexity experiments were performed (Table 1
) to identify the respective contributions of argillaceous mineralogy (Exp. I), native microbial community (Exp. II), ethanol addition (Exp. III), and microbial enrichment (Exp. IV) on the fate of the added As(V), Sb(V) and Cr(VI) under oscillating oxic-anoxic conditions.
Specifically, Exp. I was performed as an abiotic control created by sterilizing the argillaceous material at 121 °C for 15 min prior to the experiment and by adding microbial inhibitors (Na-azide, streptomycin, and kanamycin sulphates from Sigma-Aldrich) to the experimental suspension. Exp. II was performed with the argillaceous suspension without additional treatment. In Exp. III, 20 mM of dissolved organic carbon (DOC) as ethanol (≥99.8%, SIGMA-ALDRICH) was added at the beginning of each anoxic period as a source of labile OC (Wu et al., 2007). In Exp. IV, the native microbial community was enriched by adding a soil inoculum (See Appendix C), while also adding ethanol as in Exp. III. The inoculum was extracted from a natural floodplain soil located in a zone experiencing frequent redox fluctuations. Previous characterization of the microbial ecology at this field site demonstrated the presence of a broad diversity of aerobic and anaerobic bacteria (Parsons et al., 2013). Details on the preparation and injection of soil inoculum are given in Annex C.
2.3
Analyses
All standards and reagents were of analytical grade and all solutions were prepared in Milli-Q water (resistivity 18.2 MΩ cm−1). Anion concentrations (NO3
−, NO2
−, SO4
2−, Cl−, and C2H3O2
−) were determined by ion chromatography (Metrohm 761 Compact IC); total aqueous concentrations of Na, K, Ca, Mn, Fe, S, As, Cr, Sb were measured by ICP-OES (Varian 720 ES). Certified multi-elemental solutions (Sigma-Aldrich) were used for quality control with measured concentrations always within 10% of the certified values. Hereafter, we refer to aqueous concentrations using square brackets.
A UV–Vis spectrophotometer (Agilent 8453E UV–Visible) was used to measure the aqueous speciation of Cr by recording absorbance at 272 and 373 nm (Rai et al., 2007; Zydorczak et al., 2012). Speciation of aqueous Sb was measured by high performance liquid chromatography (HPLC) coupled to ICP-OES according to Lintschinger et al. (1997). Briefly, a Hamilton anion exchange column (PRP-X100, 4.1 9150 mm, 10 mm) was used with a mobile phase of 20 mM EDTA (Fluka) and 2 mM KHP (Sigma-Aldrich) (pH 4.3) at a flow rate of 1.5 ml min−1 for the separation of Sb(V) and Sb(III) peaks. Aqueous speciation of As(V) and As(III) was performed using the same column with a mobile phase of KH2PO4:K2HPO4 (Merck) at pH 6.25 prior to analysis by ICP-OES (Roig-Navarro et al., 2001).
Solid-phase concentrations of oxidized As, Sb, and Cr were determined on samples collected on the last day of each experiment. The samples were analysed by ICP-OES after leaching with a 0.1 M Na2CO3 solution (Panichev et al., 2005). The alkaline extraction is assumed to selectively release the oxidized species of As, Sb and Cr. Concentrations of reduced As, Sb, and Cr species were estimated by difference between total concentrations after acid digestion and concentrations measured in the Na2CO3 extracts (Panichev et al., 2005).
Dissolved inorganic carbon (DIC) and DOC concentrations were determined on a carbon analyser (Shimadzu TOC-Vcsn). Further characterization of DOC was performed by measuring the raw absorbance at 254 nm and calculating specific UV absorbance (SUVA 254) (Cory and McKnight, 2005). Microbial activity was inferred from temporal trends in nitrate (NO3
−) and adenosine triphosphate (ATP) concentrations (e.g., Hammes et al., 2010; Karl, 1980). The ATP concentrations were determined by measuring luminescence levels (Lumat LB 9507 Berthold Technologies) after application of a BacTiter-Glo Microbial Cell Viability Assay kit (Promega).
2.4
Biogeochemical modelling
Biogeochemical modelling was designed to identify key reactive processes and limiting factors controlling the behaviour of As, Sb and Cr during the experiments, by reproducing observed the contaminant speciation and concentration time series data. An increasing complexity modelling approach was used: parameter values derived in an earlier (simpler) experiment were fixed and used unchanged in the models for the subsequent (more complex) experiments.
Contaminant fate in the least complex experiment (Exp. I) was quantitatively assessed by fitting the multi-reaction model of Zhang and Selim (2005) to the time series of As, Sb, and Cr concentrations. The numerical solution was obtained by solving a set of nonlinear differential and algebraic equations (see detailed model description in Appendix D) in PHREEQC v.3.1.7 (Parkhurst and Appelo, 2013). The model simulates cycles of oxic and anoxic conditions by alternating the equilibrium headspace composition controlled by O2(g) (0.2 atm) and CO2(g) (0.01 atm), respectively. Equilibrium with pCO2 and calcite was imposed to buffer the pH. The WATEQ4F database was used as a starting point and additional equilibrium constants from the literature were used as required (Table D.1, D.2).
The multi-reaction model for Exp. I was run for each of the three contaminants to derive parameter values for the reversible equilibrium sorption constant (Keq), the rate constants for reversible adsorption (kf) and desorption (kb), plus the rate constant for irreversible sorption (ki). The model for Exp. II incorporated all reactions used for Exp. I, supplemented by microbially driven reduction reactions with rates formulated based on modified Monod kinetics (e.g., Van Cappellen and Wang, 1996). Chemical and microbial reactions were assumed to occur in parallel. No changes to the model of Exp. II were required to reproduce contaminant behaviour in Exp. III. Finally, the most complex model, that for Exp. IV, was obtained by adding rates accounting for microbial respiration to the model for Exp. II-III. The entire reaction network used in the biogeochemical models is given in Table 2
. The assumptions and kinetic formulations are explained in the discussion section where we refer to the reaction numbers indicated in the table.
An optimization algorithm was implemented in Python. The scipy.optimize package (Jones et al., 2001) was used to minimize root-mean-square errors (RMSE) and select the subset of parameters yielding the best fit between model and data. A global sensitivity analysis was performed to ascertain how the predictive models depend on parameter values via a Fourier Amplitude Sensitivity Test (FAST) (Saltelli et al., 1999) by using the open source Python library SALib (Herman and Usher, 2017). The sensitivity of the predicted aqueous concentrations to parameter values was analysed by varying parameters one at a time (“First”) and by accounting for the sum of the interaction effects (“Total”). The respective sensitivity indices were estimated by running the models with a distribution of parameters within a range spanning four orders of magnitude.
3
Results
3.1
Argillaceous matrix characterization
The mineralogical composition of the Tégulines argillaceous geomaterial (Table A.1) was dominated by clay minerals (muscovite, smectite, kaolinite, biotite, chlorite, microcline) (56%), quartz (29%), calcite (12%), and minor amounts (<2%) of gypsum, pyrite, and anatase. The material was depleted in reducible Fe plus Mn minerals and reactive Al phases (CDB extract), which together comprised less than 0.2% dry weight (d.w.) (Table A.2). The major cations indicated typical clay composition (Table A.2): SiO2 (56%), Al2O3 (14%), CaO (8%), Fe2O3 (5%), K2O (2%), MgO (1%), TiO2 (0.7%), Na2O (0.2%), P2O5 (0.1%). The residual dry weight (11%) was lost on ignition.
3.2
Redox potential (EH), pH and microbial activity
The imposed cycles of oxic and anoxic conditions resulted in reproducible fluctuations of EH values, which remained above +200 mV throughout all experiments (Fig. 1
: a-d). The EH of the sterilized argillaceous suspension (Exp. I) oscillated between +368 ± 8 and +338 ± 10 mV under oxic and anoxic conditions, respectively. The EH profiles were similar in Exp. II and Exp. III with average oxic and anoxic values of +345 ± 5 and +338 ± 10 mV, respectively. The corresponding EH values in Exp. IV were +342 ± 4 (oxic) and +245 ± 15 (anoxic) mV. The pH remained stable at near-neutral values (7.1 ± 0.1) throughout all experiments (Fig. 1: a-d).
In Exp. I, the NO3
− concentration remained constant at its initial level (488 ± 11 μM) and no nitrite (NO2
−) was detected in solution; ATP concentrations were low or undetectable throughout the experiment (Fig. 1: e). When trace concentrations of ATP were measured (<0.3 nM) on days 17 and 35, additional injections of microbial inhibitors were performed. In Exp. II, NO3
− decreased during anoxic periods reaching 390 μM at the end of the experiment (initial concentration was 520 μM); up to 20 μM of NO2
− was measured, although it did not persist in solution (Fig. 1: f). At the beginning of Exp. II, ATP was 0.3 nM; it increased thereafter, stabilizing at 0.7 nM during the first 19 days, then decreasing to 0.1 nM and remaining stable at this level (Fig. 1: f).
In Exp. III, with ethanol additions, NO3
− concentrations decreased at a similar rate as in Exp. II over the first 30 days while ATP remained stable at 0.20 ± 5 nM. By the end of the 3rd anoxic period, NO3
− reduction accelerated and NO2
− increased up to 395 μM during the 4th anoxic period, coincident with an increase of ATP to 1.24 nM (Fig. 1: g). During the oxic period between the 3rd and 4th anoxic periods, the concentrations of NO3
−, NO2
−, and ATP remained constant. In Exp. IV, the addition of soil inoculum resulted in the highest initial ATP concentration of 1.3 nM, and the fastest NO3
− reduction to NO2
−, followed by a complete removal of NO3
− and NO2
− from solution during the 1st anoxic period. In the absence of NO3
− and NO2
−, ATP concentrations continued to gradually increase, reaching 5.0 nM by the 4th anoxic period (Fig. 1: h). ATP concentrations decreased during oxic periods (higher EH) and increased during anoxic periods (lower EH) resulting in a statistically significant negative correlation between ATP and EH (r = −0.22, p < .005).
3.3
Solution chemistry
During the pre-equilibration period, 0.41 ± 0.04 mM (0.1 mg g−1 solid) of DOC was released from the argillaceous matrix into solution in all four experiments. The low absorbance (UVA254 = 0.2, SUVA = 0.04) suggested a primarily aliphatic structure. In the absence of ethanol additions, DOC accumulated in solution to 47.8 and 42.7 mM in Exp. I and Exp. II, respectively (Fig. B.2). The constant rate of DOC accumulation significantly correlated with that of Cl− (r = 0.99, p < .005) indicating a common source: leakage from the pH and EH electrodes (Fig. B.1). Further supporting information can be found in Appendix B.
During experiments with ethanol additions we observed a step-wise increase in total DOC concentration up to 54.3 mM (Exp. III) and 56.5 mM (Exp. IV), of which, acetate accounted for 5 and 7.5 mM, respectively (Fig. 2: g-h). In Exp. IV, 1.8 mM acetate was initially present due to the addition of the soil inoculum. It became depleted to 0.25 mM during the first 15 days and accumulated in solution up to 7 mM during the rest of the experiment. The decrease in acetate concentration was coincident with increased DIC concentration from 4 to 7 mM by day 15 (Fig. 2: h). The concentration of DIC was relatively stable (4.05 ± 0.5 mM) in all other experiments (Fig. 2: e-g).
Calcite dissolution (SI = −0.33 at 25 °C, pCO2 = 10−2 atm) likely controlled the increase of [Ca] from 3 to 6 mM in all experiments. Total [Fe] was generally ∼2 μM (Fig. B.1: a); [Mn] was around 0.5–0.9 μM initially and remained constant in Exp. I – II, while it increased to 2.7 and 5.7 μM by the end of Exp. III and Exp. IV, respectively (Fig. B.1: b); [S] remained unchanged with time and was always present as ∼6 mM sulfate (SO4
2−) (Fig. B.1: c).
3.4
Contaminant behaviour
Under abiotic conditions in Exp. I, contaminant behaviour was not affected by the oscillations between oxic and anoxic cycles. Total aqueous [As] continuously decreased from 460 to 159 μM, [Sb] from 528 to 352 μM, and [Cr] from 560 to 235 μM (Fig. 2: a). Over the course of the experiment, twice as much As and Cr were removed from solution compared to Sb. Aqueous contaminant speciation analyses indicated that all contaminants remained in their initial oxidized states (Fig. B.3). Solid-phase As and Sb speciation analyses suggested that these elements were sequestered by the argillaceous matrix in their oxidized forms. In contrast, 90% of Cr was immobilized in the solid phase as reduced Cr(III) (Fig. A.2).
In Exp. II, [As] steadily decreased from 460 to 105 μM during the first 30 days and increased only slightly afterwards to 135 μM (Fig. 2: b). The decrease of [As] was coincident with removal of As(V) from solution and the appearance of As(III) (Fig. B.3: a). Selective extraction results suggested that 83% of sequestered As in the solid-phase was present as As(V) and 17% as As(III) by day 49 (Fig. A.2: a). Slightly more aqueous Sb was removed in Exp. II than in Exp. I. This is consistent with greater solid phase accumulation of Sb in Exp. II, of which 20% occurred as Sb(III) (Fig. A.2: b). Aqueous [Cr] decreased below the detection limit (0.05 μM) by the last day of the experiment with concurrent accumulation in the solid phase as Cr(III) (>98%) (Fig. A.2: c).
In Exp. III, aqueous As, Sb, and Cr were sequestered as in Exp. II, although at slightly faster rates. [As] decreased from 460 to 110 μM over 26 days (Fig. 2: c), concomitant with complete reduction of aqueous As(V) to As(III) (Fig. B.3: a), and As(III) accumulation in the solid-phase by the end of the experiment. Antimony, again, was sequestered to a lesser extent than the other contaminants and remained in solution entirely as Sb(V), though, about 15% of solid-phase Sb was found to be Sb(III). [Cr] behaved as in Exp. II, although depletion from solution was faster, which appeared to trigger the onset of NO3
− reduction.
In Exp. IV, the contaminants were sequestered to the greatest extent. In this experiment, the removal rates were affected by the oscillating oxic and anoxic conditions (Fig. 2: d). Decreases in [As] were only observed under anoxic conditions; after day 15 all remaining aqueous As was under the form of As(III). Complete As(V) removal from solution appeared to enhance Sb(V) removal: after day 15, [Sb] decreased at twice the initial rate. By the end of the experiment, only Sb(V) was detected in the aqueous phase (Fig. B.3: b), while oxidized and reduced Sb were detected in a 4:1 ratio in the solid-phase (Fig. A.2). Chromium(VI) was completely reduced and removed from solution to below detectable concentrations (<0.05 μM) during the first anoxic period.
Comparison of contaminant removal from the aqueous phase with accumulation in the solid phase at the end of the experiments showed good mass balances (Fig. A.1). The only exception was Sb removal from solution in Exp. IV, which indicated a 18% Sb deficit from the solid-phase. Overall, sequestered contaminants were predominantly present as As(V), Sb(V), and Cr(III) comprising >75% of their total solid-phase accumulation by day 49 in Exp. IV (Fig. 5).
4
Discussion
4.1
Abiotic sequestration of As, Sb, Cr in Exp. I
4.1.1
Distribution coefficients (Kd)
We used the model derived for Exp. I to extrapolate the aqueous contaminant concentrations in equilibrium with the argillaceous mineral matrix beyond the experimental time period. The modelling results suggest that equilibrium partitioning would be reached after 72 days for Sb(V), 113 days for As(V), and 400 days for Cr(VI). The predicted contaminant equilibrium concentrations were used to calculate distribution coefficient (Kd) values (Table 3
) specific to the experimental conditions: suspension density (50 g L−1), total contaminant concentrations (∼500 μM), pH (7.1 ± 0.1), ionic strength (0.05 M), and temperature (25 °C).
Derived Kd values for As and Sb are lower than those reported in the literature for pure clay phases. For example, at near-neutral pH, Kd values of Sb(V) in pure kaolinite suspensions vary from 73 to 115 L kg−1 according to the experimental data of Rakshit et al. (2015). This Kd range can be reconciled with the Kd value (9.8 L kg−1) calculated for Sb(V) in the present study when taking into account that kaolinite represents 11% d.w. of the argillaceous matrix. Similarly, the Kd value of As(V) in Exp. I (46.7 L kg−1) is consistent with the Kd for pure muscovite at neutral pH (282–416 L kg−1) (Chakraborty et al., 2007), given the 15% d.w. content of muscovite in the argillaceous matrix. The Kd value of Cr is much higher than those of As and Sb because of the inferred reductive precipitation mechanism, which completely removes Cr from solution.
4.1.2
Arsenic abiotic sequestration
Similar to As(V) sequestration by pure mineral phases (Couture et al., 2013) and natural soils (Zhang and Selim, 2005), As(V) adsorption to the Tégulines matrix in Exp. I is best described by a two-phase process with an initial rapid (equilibrium) uptake followed by a slow kinetic immobilization (Fig. 3
: a). The model yields an apparent equilibrium sorption constant log K = −0.54 (Table 3), which describes the instantaneous removal of 30% of the aqueous As(V), likely via the formation of outer-sphere complexes. The reversible kinetic uptake, presumably through the formation of inner-sphere complexes, is represented by the adsorption (kf) and desorption (kb) rate constants. The kf and kb values for As(V) in this study are up to 3 orders of magnitude lower than those estimated for As(V) sorption on Fe-bearing minerals (2-line ferrihydrite, goethite, amorphous mackinawite, pyrite; Couture et al., 2013) and natural soils containing 5.7 ± 2 mg g−1 of amorphous Fe2O3 (Oliver loam, Sharkey clay, Windsor sand; Zhang and Selim, 2005).
Clays offer a variety of sites suitable for As(V) sorption on (1) broken clay edges via surface ligand exchange (Charlet et al., 2012; Lin and Puls, 2000; Wilson et al., 2010), (2) negatively charged clay surfaces (e.g., kaolinite) at neutral pH via cation (e.g., Ca2+) bridging (Sharma and Kappler, 2011), and (3) hydroxide interlayers in the clay structure of chlorite (Lin and Puls, 2000). All of these mechanisms are likely to play a role in our experiments, given the presence of kaolinite (11% d.w.) and chlorite (3.9% d.w.), pH of 7.2 and high [Ca] (4 mM). Sorption onto calcite (12% d.w.) at ambient carbon dioxide pressure (pCO2 = 10−3.5 atm) would require a higher pH (>8.3) (Sø et al., 2008).
Slow As(V) removal, which is implemented in the model as an irreversible kinetic reaction, most probably represents intra-/inter-particle diffusion (Zhang and Stanforth, 2005) or the formation of surface precipitates (Gallegos et al., 2007), such as Ca-arsenate phases (Bothe and Brown, 1999; Raposo et al., 2004). Based on the initial [As] concentration in Exp. I, the experimental solution is oversaturated with respect to several Ca-As phases: Ca3(AsO4)2(H2O)4 (SI = 6.1) and Ca3(AsO4)2 (SI = 6.3). Hence, thermodynamically, the formation of such precipitates is possible, although likely rate-limited.
4.1.3
Antimony abiotic sequestration
Modelling results suggest that sequestration of Sb(V) (Fig. 3: e) is mainly kinetically-controlled with the initial equilibrium uptake accounting for only 5% of total Sb sequestration (log K = −1.07) (Table 3). This is consistent with results suggesting low Sb(V) sorption on clay minerals (i.e., kaolinite) at neutral pH (Rakshit et al., 2015). Estimated rate constants of Sb(V) adsorption and desorption are on the same order of magnitude as those of As(V). The fact that the Sb(V) desorption rate constant (kb) is higher than that of sorption (kf), together with the very low irreversible sequestration rate constant (ki), suggests that Sb(V) uptake is mostly reversible, as previously observed (Johnson et al., 2005; Xi et al., 2010), although precipitation of Ca[Sb(OH)6]2 (Johnson et al., 2005) is thermodynamically possible (SI = 3.33) in Exp. I.
4.1.4
Chromium abiotic sequestration
In contrast to As(V) and Sb(V), which remained oxidized under abiotic conditions, Cr(VI) was reduced and accumulated in the solid phase as Cr(III). Chromium is a strong oxidant which can be reduced at surface sites of Fe(II)-bearing clay minerals and pyrite (Chon et al., 2006) to form Cr(OH)3(am), Cr(OH)3⋅H2O(cr), Cr2O3, or (Cr,Fe)(OH)3 (Nriagu et al., 1993), even in the presence of O2 (Rai et al., 1989). The continuous decrease of [Cr] during oxic and anoxic periods (Fig. 3: i) corroborates that abiotic Cr reduction is not inhibited by O2. Moreover, Cr(VI) can be reduced by DOC (Jamieson-Hanes et al., 2012) released from the argillaceous matrix itself.
The measured speciation of solid-phase Cr implies a two-step uptake mechanism with adsorption of Cr(VI) followed by its abiotic reduction leading to Cr sequestration as Cr(OH)3 (SI = 1.18) (Fig. 3: i). The modelled mechanism is consistent with the XANES results of Jamieson-Hanes et al. (2012) demonstrating that Cr reduction occurs at particle surfaces rather than in solution. Assuming constant pH, we introduce a first-order kinetic reduction rate of adsorbed Cr(VI) (Table 2, R28). The rate constant derived from fitting the data (3.9 × 10−3 h−1; Table D.2) is close to that reported for soil (6.9 × 10−3 h−1; Matern and Mansfeldt, 2016). Overall, the model suggests weak Cr(VI) sorption on the argillaceous matrix consistent with previous reports for kaolinite, montmorillonite, and illite (Lan et al., 2008). Abiotic sequestration of Cr(VI) is thus interpreted as mainly resulting from reductive precipitation, which sequesters ∼50% of aqueous Cr with the rest (∼8%) being removed via surface adsorption (Fig. 5).
4.2
Microbial effects on the mobility of As, Sb, Cr in Exp. II, III, IV
4.2.1
Microbial activity
The observed temporal trends of NO3
− and ATP concentrations indicate that in Exp. II – III microorganisms were present and active in the bioreactors, while they were inactive in the sterilized suspension of Exp. I (Fig. 1: e-h). Nitrate reduction is associated with energy production (Konhauser, 2007) and a corresponding ATP increase is observed in Exp. III (Fig. 1: k): ATP concentrations and NO3
− (Exp. III – IV) reduction rates are significantly correlated (r = 0.99, p < .001).
The electron donor for NO3
− reduction is likely naturally present in the argillaceous matrix. Ethanol is an unlikely reductant, as the NO3
− reduction rate did not respond to the ethanol additions during the 1st, 2nd, and 3rd anoxic periods of Exp. III. Theoretically, DOC released by the argillaceous matrix during the pre-equilibration period (0.1 mg g−1) may attenuate about 12 μmol g−1 of NO3
−, assuming a zero oxidation state of DOC. Moreover, pyrite within the argillaceous matrix may attenuate up to 417 μmol g−1 of NO3
−. Even though nitrate reduction coupled to organic carbon and pyrite oxidation are both widespread processes under biotic conditions (Bosch et al., 2012), the higher energy yield of NO3
− reduction with organic matter should favor this reaction pathway over anaerobic pyrite oxidation (Appelo and Postma, 2005). Complete reduction of NO3
− by pyrite would also lead to an increase in SO4
2− and Fe by 350 and 175 μM, respectively, which was not observed (Fig. B.1: c). We therefore model nitrate consumption as a heterotrophic denitrification pathway with complete NO3
− turnover via NO2
− to N2(g) coupled to the oxidation of DOC (Table 2, R32,33) (Rivett et al., 2008).
4.2.2
Fate of As in the presence of microorganisms
The occurrence of As(V) reduction in Exp. II – III, which did not occur in Exp. I (Fig. A.2, Fig. B.3), suggests a microbially mediated reaction. Reduction rates are similar regardless of the addition of ethanol, again implying that this carbon source is not used as a primary electron donor for As reduction. Microbial As reduction can either proceed via enzymatic (respiration and detoxification) or non-enzymatic (oxidation of microbially produced glutathione) pathways (Scott et al., 1993). Given that our experimental [As] values exceed the threshold of 100 μM As for expression of detoxification genes by bacteria Shewanella sp. (Saltikov et al., 2005) by 4-fold, reduction by detoxification is a plausible pathway.
The detoxification mechanism is also supported by the lack of ATP increase throughout Exp. II – III (Fig. 1: f-g), suggesting that As reduction is not an energy gaining process for the native microbial community (Lloyd and Oremland, 2006; Oremland et al., 2009; Plant et al., 2005). In agreement with the observations of Saltikov et al. (2005) who reported aerobic As(V) detoxification by Shewanella sp. ANA-3, As reduction is not inhibited by the presence of O2 in our study (Fig. 3: b-c). Moreover, other terminal electron acceptors (TEAs) of higher redox potential (Table 4
) did not inhibit the reduction of As(V) to As(III) (Fig. B.3), which occurred simultaneously with CrO4
2− and NO3
− reduction. The often-hypothesized thermodynamic cascade of TEA utilization is therefore not observed. Finally, the fact that ethanol additions did not increase the As reduction rate supports detoxification over respiration as an As-reduction mechanism. In the absence of a suitable external electron donor, possible electron donors for detoxification are endogenous reservoirs (e.g., DNA and proteins) mediated by intracellular reductants (e.g., NADH, NADPH, ascorbate, glutaredoxin, glutathione; Silver and Phung, 2005).
To account for both mineralogical and microbial controls on As sequestration in Exp. II – III, Monod-type kinetic expressions are added to the multi-reaction model of Exp. I. We assume that the rate of detoxification is independent of the DOC concentration and not inhibited by O2 (Table 2, R10). Modelling results indicate that the persistence of As(III) in solution is consistent with its low sorption affinity to clay minerals (Table 3; Lin and Puls, 2000) and the under-saturation of the solution with respect to CaHAsO3 (SI = −0.38). We note that the model fails to capture the observed slight increase of aqueous [As] between day ∼30 and the end of Exp. II – III (Fig. 3: b-c). Late As remobilization could reflect the re-equilibration of surface-bound As(III) with the solution under evolving geochemical conditions. Alternatively, the release of As(III) in solution could be a result of late reduction of surface-bound and, thus, less bioavailable As(V) (Ghorbanzadeh et al., 2015).
In Exp. IV, the inoculation of soil microorganisms resulted in the inhibition of As(V) reduction by O2 during the first oxic period (Fig. 3: d). Furthermore, As(V) reduction was associated with a 1:1 conversion of carbon in a form of acetate to DIC (Fig. 2: h), consistent with acetate oxidation to bicarbonate. Because acetate may provide electrons for microbial As reduction (Lloyd and Oremland, 2006; Oremland et al., 2009), it can enable As respiration (Gorny et al., 2015). An increase in ATP concentration, expected for enhanced microbial respiration (Oremland, 2005), is also observed (Fig. 1: h).
In the model, As(V) respiration is simulated by a rate expression that accounts for inhibition by O2 and depends on the availability of organic carbon (acetate) (Table 2, R11). Initial model runs, however, overestimated [As] remaining in solution and underestimated As(V) accumulation in the solid phase (Fig. D.3: b). A potential mechanism, accounting for the missing As(V), is sorption of non-volatile mono-methyl As (MMAs(V); Lafferty and Loeppert, 2005) formed via oxidative biomethylation of As(III) (Mestrot et al., 2013). This mechanism is speculative due to the lack of direct experimental confirmation, but weakly supported by its inhibition at the initiation of Sb(V) reduction. The inhibition of As biomethylation in the presence of Sb(III) has been demonstrated by Andrewes et al. (2000). We therefore added the oxidative biomethylation of As(III) in the model (Table 2, R16). The modelling exercise reveals that ∼46% of As(III) could be immobilized via sorption of freshly formed methylated As(V) (Fig. 3: d). The overall aqueous As(V) concentration predicted by the model is most sensitive to the rates of abiotic adsorption (
k
f
H
A
s
O
4
2
−
) and microbial respiration (
k
r
e
s
p
H
A
s
O
4
2
−
) (Fig. D.4) suggesting that both sorption and reduction mechanisms play significant roles in As(V) sequestration (Fig. 5).
4.2.3
Fate of Sb in the presence of microorganisms
Similar to As, reduction of Sb(V) to Sb(III) is mediated by native microorganisms and suggestive of a detoxification mechanism in Exp. II – III (Fig. 3: f-g). Based on the saturation index, precipitation of Sb(III) in the form of Sb2O3(s) (SI = 0.69) is thermodynamically possible, consistent with microbial Sb(V) reduction (Abin and Hollibaugh, 2014). Fast reductive precipitation of Sb agrees with the absence of aqueous Sb(III) and the co-existence of Sb(V) with As(III) in the experimental solutions of Exp. II and III, as previously demonstrated in other studies (Casiot et al., 2007; Fawcett et al., 2015; Mitsunobu et al., 2006). The model fit to the experimental data requires a two-step process: sorption of Sb(V) followed by its reduction at the mineral surfaces, presumably via a detoxification pathway (Table 2, R21). This would suggest that the microorganisms capable of Sb detoxification are surface-associated rather than planktonic. To our knowledge, such a mechanism has not yet been described in the literature.
In Exp. IV, after complete As(V) removal by day 15, the removal rate of aqueous Sb increased, but only during the anoxic periods (Fig. 3: h). Recent studies have demonstrated that under anoxic conditions, Sb(V) can act as a TEA for microbial respiration (Abin and Hollibaugh, 2014), but also that Sb(V) respiration is inhibited by the presence of aqueous As(V) (Fawcett et al., 2015; Kulp et al., 2014; Mitsunobu et al., 2006). Thus, Sb(V) respiration could explain the increase of ATP observed in Exp. IV following the complete reduction of alternative TEAs (CrO4
2−, NO3
−, and HAsO4
2−) (Fig. 1: h).
The observed sequence of As(V) reduction prior to Sb(V) reduction agrees with the thermodynamic sequence based on the calculated EH values of the TEA reduction reactions (Table 4). Under the experimental conditions (pH ∼7, 25 °C), As(V) is expected to suppress Sb(V) reduction to solid-phase Sb2O3(s), but not to aqueous Sb(OH)3(aq). Microbial reductive precipitation of Sb2O3(s), coupled to the oxidation of lactate to acetate, has been reported by Abin and Hollibaugh (2014). Likewise, in Exp. IV, a gradual accumulation of acetate in solution after day 15 (Fig. 2: h) suggests that Sb reduction is coupled to the oxidation of ethanol to acetate. Although a large number of studies have previously compared the biogeochemical properties of Sb and As (e.g., Dovick et al., 2016; Willis et al., 2011; Wilson et al., 2010), the incomplete oxidation of organic carbon by Sb(V), in contrast to complete mineralization to DIC by As(V), has not previously been highlighted.
Inclusion of Sb(V) reduction in the model, via a respiration mechanism which is inhibited by dissolved O2 and aqueous As(V) (Table 2, R22), reproduces the features of the Sb time-series observed in Exp. IV (Fig. 3: h). However, the failure of the model to account for Sb(III) accumulation in the solid phase is possibly related to the experimentally observed mismatch between Sb loss from solution and that accumulated in the solid-phase (18%) (Fig. A.1). In the presence of As(III) and As(V), the missing Sb could be volatilized via methylation (Hartmann et al., 2003). Overall, sensitivity analyses identify the rate of adsorption (
k
f
S
b
(
O
H
)
6
−
) as the most critical parameter controlling the mobility of Sb(V) in the argillaceous suspension (Fig. D.4).
4.2.4
Fate of Cr in the presence of microorganisms
Sequestration of Cr in the argillaceous suspensions likely proceeds via reductive precipitation of Cr(OH)3(s), which is enhanced in the presence of microorganisms in Exp. II – III as compared to abiotic Exp. I (Fig. 3: i-k). Similar to As and Sb, native microorganisms could induce Cr(VI) reduction via detoxification (Cheung and Gu, 2007). Reductive detoxification of both Cr and Sb may reflect a microbial strategy for metal sequestration via precipitation. Alternatively, Cr(VI) could be reduced via extracellular non-enzymatic processes. In this case bacteria could serve as electron donors after biosorption of Cr to the microbial cell wall (Viti et al., 2014). To account for the microbial control on Cr sequestration, we introduce a microbial pathway of Cr(VI) reduction to the model (Table 2, R29).
In Exp. IV, rapid Cr removal during the first 3 days (Fig. 3: l) likely reflects the addition of soil inoculum capable of Cr respiration (Viti et al., 2014). We therefore, include Cr respiration (Chen and Hao, 1998) into the model (Table 2, R30). Our estimate for the rate constant of reduction falls within the range reported for Cr(VI) reduction by dissimilatory iron-reducing bacteria (0.8 × 10−4 – 3.1 × 10−4 h−1) (Table D.2) (Fendorf et al., 2000). Note that although some removal of aqueous Cr by mineral particles in the added soil inoculum cannot be excluded, we do not believe this is an important process. Sensitive model parameters include the maximum rate of detoxification (
r
m
a
x
C
r
O
4
2
−
), as well as the rate constants of abiotic reduction (
k
a
b
C
r
O
4
2
−
), sorption (
k
f
C
r
O
4
2
−
), and desorption (
k
b
C
r
O
4
2
−
) (Fig. D.4). Consistent with the experimental observations, the model-predicted temporal changes of the aqueous Cr(VI) concentration is relatively insensitive to the value imposed for the O2 inhibition constant (
K
O
2
i
n
).
4.3
Respiration versus detoxification
The experimental results provide evidence that the fate of the contaminant oxyanions (i.e., HAsO4
2−, Sb(OH)6
−, and CrO4
2−) added to the argillaceous suspensions depends on the level of microbial activity (as inferred from NO3
− and ATP data) and the microbial diversity and functionality (native versus inoculated consortia). The experimental results further suggest two distinct strategies employed by the microorganisms to deal with the mixed contaminants (Fig. 4
).
The first strategy involves the simultaneous reduction of HAsO4
2−, Sb(OH)6
−, and CrO4
2− under aerobic and denitrifying conditions. This is observed in the suspensions of Exp. II – III (Fig. 2: b-c). It is consistent with the previously observed simultaneous reduction of SO4
2− and As(V) (Macy et al., 2000) or Fe(III) and As(V) (Smeaton et al., 2012), in which As is reduced as part of a detoxification pathway. Furthermore, the suggested detoxification mechanism may explain the bioreduction of other potentially toxic elements, for example U(VI), which can be reduced alongside of Fe(III) or SO4
2− in sediments (Finneran et al., 2002; Komlos et al., 2008).
The second strategy involves the respiratory reduction of the oxyanions that follows the thermodynamic sequence from highest to lowest reduction potential (Table 4). Such a sequential reduction order is observed when the argillaceous suspensions are enriched with the soil inoculum in Exp. IV (Fig. 2: d). The corresponding sequence of TEA utilization in Exp. IV is then: O2, CrO4
2−, NO3
−, HAsO4
2− and Sb(OH)6
−.
4.4
Redox potential of the argillaceous suspension
Measured EH values of the argillaceous suspensions during anoxic periods (+283 ± 65 mV) do not correspond to the values predicted by the Nernst equation for the experimental conditions based on the redox couples: CrO4
2−/Cr(OH)3, NO3
−/NO2
−, HAsO4
2−/H3AsO3 and Sb(OH)6
−/Sb2O3(s) (Table 4). Even though the activity ratios of the redox couples imply a range of redox conditions, from oxidizing (Exp. I) to reducing (Exp. II – IV), the measured EH values do not respond quantitatively to these conditions and remain systematically above +200 mV. Therefore, the measurement of EH does not yield a quantitative indicator of Cr, N, As, or Sb redox processes, most likely because none of the corresponding redox couples are electroactive towards the redox electrode (Markelova et al., 2017). Overall, our results confirm that EH measurements are not particularly informative in environments depleted in electroactive redox couples (e.g., Fe couples), such as argillaceous formations.
4.5
Stability of reduced contaminants
We did not observe re-oxidation of reduced As(III), Sb(III) and Cr(III) during the oxic periods, even though oxidation of the reduced forms of these elements by O2 is thermodynamically favorable under the experimental conditions (Campbell and Nordstrom, 2014). The apparent stability of the reduced species observed here contrasts with previous studies of contaminant dynamics in topsoil suspensions under similar redox-oscillating conditions (e.g., Couture et al., 2015; Hockmann et al., 2014; Parsons et al., 2013), where re-oxidation is observed to various degrees. We hypothesize that the lack of re-oxidation reflects kinetic limitation in argillaceous materials lacking reactive Fe- and Mn-oxyhydroxides (Table A.2) (Leuz and Johnson, 2005; Rai et al., 1989). It is also possible that oxidizing microorganisms were not fully established in the reactor system due to the short experimental incubation period. These factors have all been previously linked to the re-oxidation of contaminants in soils (Buschmann et al., 2005; Butler et al., 2015; Hug and Leupin, 2003).
5
Conclusions
The fate of oxyanions of As(V), Sb(V) and Cr(VI) in Tégulines argillaceous suspensions was monitored under oscillating oxic-anoxic conditions in an increasing complexity experimental design using a bioreactor system. A dynamic modelling approach was used to reproduce the data and provide an internally consistent interpretation of the possible pathways responsible for the sequestration of the aqueous contaminants (Fig. 5, Table D.3). The modelling results suggest that a combination of adsorption, microbial detoxification and respiration processes explains the most extensive removal of As(V) from solution observed in Exp. IV. Moreover, we speculate that in Exp. IV, microbial methylation of As(III) to MMAs(V) may have prevented the accumulation of toxic As(III) in solution. In contrast to As, sequestration of aqueous Sb and Cr is mainly controlled by reductive precipitation, likely forming Sb2O3(s) and Cr(OH)3(s). The reductive precipitation of Sb(III) appears to be mainly microbially mediated, contrary to Cr for which abiotic reductive precipitation is identified as the major sequestration pathway under all investigated conditions. We further propose that the ability of the native microbial community to simultaneously reduce oxidized As, Sb and Cr, irrespective of the thermodynamically predicted order and independent of the availability of dissolved organic carbon, may be diagnostic of contaminant detoxification (Exp. II – III). In contrast, a sequence of contaminant reduction reactions that follows the order of decreasing redox potentials of electron acceptors (in Exp. IV: O2, CrO4
2−, NO3
−, HAsO4
2−, Sb(OH)6
−) is suggestive of respiration rather than detoxification. Keeping in mind that we only considered short-termed oxidative perturbations (7 days), our experimental and modelling results indicate that the investigated argillaceous formation may constitute a suitable geochemical environment for contaminant sequestration, as exposure to O2 did not cause the re-release of the contaminants to the aqueous phase. This conclusion, however, will have to be confirmed for redox oscillations lasting on longer time scales (e.g., several months and longer).
Acknowledgments
We acknowledge Delphine Tisserand (University of Georgia) and Marianne Vandergriendt (University of Waterloo) for support in the laboratory. We are grateful to Nele Bleyen and Hugo Moors from SCK-CEN for helpful discussions on the geomicrobiology of the argillaceous formations. We appreciate a detailed feedback of Grant Ferris to this work reviewed as part of EM's PhD thesis. This research was funded by Andra (the French National Radioactive Waste Management Agency), the Geochemistry group at ISTerre, which is a part of Labex OSUG@2020 (ANR 10 LABX 56), and the Canada Excellence Research Chair program.
Appendix A
Supplementary data
The following is the supplementary data related to this article:
Annex_final
Annex_final
Appendix A
Supplementary data
Supplementary data related to this article can be found at https://doi.org/10.1016/j.apgeochem.2017.12.012.
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16537
16543
10.1021/ie302096e
AG
4007
S0883-2927(17)30376-1
10.1016/j.apgeochem.2017.12.012
The Authors
Speciation dynamics of oxyanion contaminants (As, Sb, Cr) in argillaceous suspensions during oxic-anoxic cycles
Ekaterina
Markelova
a
b
∗
emarkelo@uwaterloo.ca
ekaterina.markelova@ujf-grenoble.fr
Raoul-Marie
Couture
b
c
rmc@niva.no
Christopher T.
Parsons
b
chris.parsons@uwaterloo.ca
Igor
Markelov
a
b
igor.markelov@uwaterloo.ca
Benoit
Madé
d
Benoit.Made@andra.fr
Philippe
Van Cappellen
b
pvc@uwaterloo.ca
Laurent
Charlet
a
b
laurent.charlet@ujf-grenoble.fr
a
University Grenoble Alpes, University Savoie Mont Blanc, CNRS, IRD, IFSTTAR, ISTerre, 38000 Grenoble, France
University Grenoble Alpes
University Savoie Mont Blanc
CNRS
IRD
IFSTTAR
ISTerre
Grenoble
38000
France
b
Ecohydrology Research Group, Department of Earth and Environmental Sciences and Water Institute, University of Waterloo, 200 University Ave. W, Waterloo, ON N2L 3G1, Canada
Ecohydrology Research Group
Department of Earth and Environmental Sciences and Water Institute
University of Waterloo
200 University Ave. W
Waterloo
ON N2L 3G1
Canada
c
Norwegian Institute for Water Research-NIVA, Gaustadalléen 21, 0349 Oslo, Norway
Norwegian Institute for Water Research-NIVA
Gaustadalléen 21
Oslo
0349
Norway
d
Andra, National Radioactive Waste Management Agency, R&D Division, Transfer Migration Group, 1/7 rue Jean Monnet, 92298 Chatenay Malabry Cedex, France
Andra
National Radioactive Waste Management Agency
R&D Division
Transfer Migration Group
Chatenay Malabry Cedex
1/7 rue Jean Monnet
92298
France
∗
Corresponding author. University Grenoble Alpes, University Savoie Mont Blanc, CNRS, IRD, IFSTTAR, ISTerre, 38000 Grenoble, France.
University Grenoble Alpes
University Savoie Mont Blanc
CNRS
IRD
IFSTTAR
ISTerre
Grenoble
38000
France
Abstract
Argillaceous geological formations are considered promising repositories for waste containing inorganic contaminants. However, the sequestration capacity of an argillaceous natural barrier may change as a result of dynamic environmental conditions, in particular changes in redox state. Here, we imposed redox cycles to argillaceous suspensions amended with a mixture of the inorganic contaminants As(V), Sb(V), Cr(VI), by alternating 7-day cycles of sparging with either oxic or anoxic gas mixtures. During the redox cycles, we assessed the relative importance of different contaminant sequestration mechanisms under sterile and non-sterile conditions. The non-sterile experiments were carried out 1) with the argillaceous material as is, that is, containing the native microbial population, 2) with the addition of ethanol as a source of labile organic carbon, and 3) with the addition of both ethanol and a soil microbial inoculum. In order to capture the observed solid-solution partitioning trends, a series of mixed thermodynamic-kinetic models representing the dynamic conditions in the experiments were developed using PHREEQC. Parameter values for processes such as adsorption, which occurred in all experiments, were kept constant across all the simulations. Additional formulations were introduced as needed to account for the microbial detoxification and respiration of the contaminants in the increasing complexity experiments. Abiotic adsorption of As and abiotic reductive precipitation of Cr dominated sequestration of these two contaminants under all experimental conditions, whereas the presence of microorganisms was essential to decrease the aqueous Sb concentration. Once reduced, As(III), Sb(III), and Cr(III) were not re-oxidized upon exposure to dissolved O2 in any of the experiments. Neither the mineralogy nor the native microbiota catalysed contaminant oxidation processes. Thus, kinetically, the reduction of As(V), Sb(V), and Cr(VI) was irreversible over the experimental duration (49 days). The capacity of the argillaceous matrix to reduce and sequester the aqueous contaminants, without subsequent remobilization to the aqueous phase during the oxic periods, supports the use of argillaceous barriers for geologic waste storage.
Graphical abstract
Image 1
Highlights
•
Aqueous As(V) is removed by adsorption, Cr(VI) by abiotic reductive precipitation.
•
Microorganisms remove aqueous Sb(V) via reductive precipitation.
•
Once reduced, As(III), Sb(III) and Cr(III) are not oxidized when exposed to O2.
•
Oxyanions are reduced in parallel via detoxification, sequentially via respiration.
•
Measured EH does not respond quantitatively to observed speciation of Cr, N, As, Sb.
Keywords
Argillaceous suspension
Nuclear waste repository
Redox oscillations
Arsenic
Antimony
Chromium
Sequestration
Reduction
Respiration
Detoxification
KBJ00000000010497
2019-04-10T18:56:06
S300.2
S300
S0883-2927(17)30376-1
10.1016/j.apgeochem.2017.12.012
AG
0883-2927
4007
FLA
NON-CRC
UNLIMITED
NONE
2017-12-15T16:24:57Z
08832927/v91sC/S0883292717303761/main.xml
263857
MAIN
JA 5.5.0 ARTICLE
FULL-TEXT
08832927/v91sC/S0883292717303761/main.assets/fx1.sml
14456
IMAGE-THUMBNAIL
08832927/v91sC/S0883292717303761/main.assets/gr1.sml
17709
IMAGE-THUMBNAIL
08832927/v91sC/S0883292717303761/main.assets/gr2.sml
20615
IMAGE-THUMBNAIL
08832927/v91sC/S0883292717303761/main.assets/gr3.sml
25539
IMAGE-THUMBNAIL
08832927/v91sC/S0883292717303761/main.assets/gr4.sml
14495
IMAGE-THUMBNAIL
08832927/v91sC/S0883292717303761/main.assets/gr5.sml
21796
IMAGE-THUMBNAIL
08832927/v91sC/S0883292717303761/main.assets/fx1.jpg
41399
IMAGE-DOWNSAMPLED
08832927/v91sC/S0883292717303761/main.assets/gr1.jpg
82262
IMAGE-DOWNSAMPLED
08832927/v91sC/S0883292717303761/main.assets/gr2.jpg
99849
IMAGE-DOWNSAMPLED
08832927/v91sC/S0883292717303761/main.assets/gr3.jpg
148694
IMAGE-DOWNSAMPLED
08832927/v91sC/S0883292717303761/main.assets/gr4.jpg
35453
IMAGE-DOWNSAMPLED
08832927/v91sC/S0883292717303761/main.assets/gr5.jpg
117946
IMAGE-DOWNSAMPLED
08832927/v91sC/S0883292717303761/main.assets/mmc1.pdf
1497160
APPLICATION
08832927/v91sC/S0883292717303761/main.pdf
1110401
MAIN
1.7 6.5
DISTILLED OPTIMIZED BOOKMARKED
08832927/v91sC/S0883292717303761/main.raw
89552
S0883-2927(18)X0003-1
AG
0883-2927
91
C
201804
1
228
S0883-2927(17)30376-1
10.1016/j.apgeochem.2017.12.012
75
88
main.pdf
PDF
1.7
local
1554970137856
collectiebehoudsniveau 1
2019-04-11T09:55:36.754+02:00
local
1554970138368
0
SHA-512
a675b8ce3b1d5d242d41af52504513ef05f801d8098bfa0fac8a4ef26ae35190ea1bc26217b1e67ba0a995194b03473986e53566eda6fdb4ca6af571175b27ef
java.security.MessageDigest
1110401
Adobe Acrobat Document
1.7
DIAS
62
DIAS tentative identification
main.pdf
local
1554970138369
0
SHA-512
43400adabc75fa076aee5a2828757f663e15c3c82d674898a2bfe1ef87027e9bd6df4f42d12e64dec66969496a67810596a9f903e3a7783882b682059401273b
java.security.MessageDigest
89552
not checked
main.raw
local
1554970138370
0
SHA-512
2c08233aca9687ac17d1c033b97f58a61c9a1674beefc7782a0e9afb60d72a3c60dc87cbe5b9088f5b5e4e77bbf219f234c5e4df5c71b47ef3332ec4dc297cef
java.security.MessageDigest
263857
not checked
main.xml
local
1554970138371
0
SHA-512
39df6ca318425eb08f75d1a380e63e24654568b4bdc1304be6593f44666d73dc26ee58484f9284cd03d454296dc0620e5bf70b708459d21391f847b5abc26d94
java.security.MessageDigest
2873
not checked
si1.gif
local
1554970138372
0
SHA-512
b4f15be3029b899b6dc6d00d89b72ceb07078ad1a597f008489d20d3a26996804f7d9247deacc941bcee0dcb10391565870257fb1d5c35ffe1acbfc7676706d3
java.security.MessageDigest
3568
not checked
si10.gif
local
1554970138373
0
SHA-512
0dd4d0c2714d1b719eb6d3da46f0892e6fcb924590ce759c6951eccf7a04a5f32792d51c2e68f7d703696cf17fad07d985470f51ed1f92f6bcfca6d2a284231c
java.security.MessageDigest
3593
not checked
si11.gif
local
1554970138374
0
SHA-512
fed400b71e49032dafb88ac6f44c6477f54b16a247828dff546b9bbdc132abeea0b797013294f9b9ab7b961ea25d72db7fba1db91eaa93c632e7b9987151247a
java.security.MessageDigest
3549
not checked
si12.gif
local
1554970138375
0
SHA-512
432b16ad52e5b72f7f371097cd900f8c4d553149f114a932472ef09ced1dab49b98b1a5d72ea73c568d09478e47b32dd32cec7df23de0df1a9f33269986f0d66
java.security.MessageDigest
3548
not checked
si13.gif
local
1554970138376
0
SHA-512
66c1674e4d388445d52a936f5bb712b496594eefd3c83ebccf41b7437a0b35444e604681e94c1513dc6a88bc2a8bb426c098be5c0ddcb7bb1c07f798d0cb9b12
java.security.MessageDigest
2938
not checked
si14.gif
local
1554970138377
0
SHA-512
98fd3091e501bdc45ddb3da37d0c4b63ab28ac470a195468fd07f02f4ec94bfec7d705a600d4e2f732e017870156974f08da16dae157702a2ae56143759cc9c4
java.security.MessageDigest
3164
not checked
si15.gif
local
1554970138378
0
SHA-512
4a4e9c61db962d2a28be5be8d2f547eed216d8eb696f4db0b86d24a121c8c9e2bd7fa369b6fbc67e4c592e53651ab35e3a1da3b154631f531ebf61fca0f31ea2
java.security.MessageDigest
3145
not checked
si16.gif
local
1554970138379
0
SHA-512
7e6b0e0c61ef93c41f1fb45b06ad8ee125a9b11c92f3d53af09b610f8e06bce252544e30060896d84e348e9cbfd49d23e88775ea571c88b6eae9524a0d823254
java.security.MessageDigest
3152
not checked
si17.gif
local
1554970138380
0
SHA-512
8ffee2d5fcb5b396f9fca5ae70994a0d8611807128036bed0d11e407818c672ed675cf7c0801a611ea75b531daadb3b961cd3738b442e1a8456dfe5211afae79
java.security.MessageDigest
3900
not checked
si18.gif
local
1554970138381
0
SHA-512
5012cb79c7416f029acddf703b914de44bcf503d871373b3ceb0b21ae6e20cec84410d6f7480198a622138a0bafa143de8cdcd12a1cfbff43aeba5c87bba11d5
java.security.MessageDigest
4677
not checked
si19.gif
local
1554970138382
0
SHA-512
85e87ffb8cd48f7b3fcb89224545a736fd8223b0b39df0ca0dfb186e64ea7a7bdc45301b4e37feb2e74bfa17de71792d96b66f54563d01c2627104d3799120c2
java.security.MessageDigest
2870
not checked
si2.gif
local
1554970138383
0
SHA-512
be5622ef9da87164242102a0850ce0a2762e72cc1688116767a30d766d798f77566773ad763d50a5d6703969dde85767450ba83a6e4c2f1c225b8f844dc183c2
java.security.MessageDigest
2923
not checked
si20.gif
local
1554970138384
0
SHA-512
55ed6a653f3c29a51f6bf08b3639e78a4ecbd10e5cf28b3d695708b32db2d96be00545f266c2848ccd15016d5778569120808401460c4b86c792a627fba98fb4
java.security.MessageDigest
3119
not checked
si21.gif
local
1554970138385
0
SHA-512
a8a27a994f130dbeb73faf45ac6245bc470959abbe019dc5544519c3f7d2decfac3cee486b8ad2c98097a487529e09afe8f02f838494603f047d0304154dc590
java.security.MessageDigest
3117
not checked
si22.gif
local
1554970138386
0
SHA-512
cc774a106c6e9c408e2cdb03beeb623d22b873d86e40fa25fa4aaa009faa81f7e8014fb4eba8a843aaf2ca9e05ca6f595b2740eb3051098c631c5beb334e1615
java.security.MessageDigest
3135
not checked
si23.gif
local
1554970138387
0
SHA-512
70c9c74d953fe5b6756ca7776ceea95438830a3ae3b9c833ff14cd6a42f30995e7bc05c69dcefaac35c1c83da46ee0c6adf94371a0af1c8b3d808739b8ff6c18
java.security.MessageDigest
3940
not checked
si24.gif
local
1554970138388
0
SHA-512
134e8bd1edf2c14c3244d5e9b4ed1526438a0651febec620649d51ba9d9f652713369f9a663cc3d7b011c5070a7326bcbc80795b1e33e05320888ec834fd632d
java.security.MessageDigest
2984
not checked
si25.gif
local
1554970138389
0
SHA-512
326e2c17d63286f864390dc038168821ab29e1c37d04a99ab6a50ddeff93417478e60e436722e4e270afc8eb5c3ea5754d4f5e3a2a12772cdd852efa2396de5f
java.security.MessageDigest
3262
not checked
si26.gif
local
1554970138390
0
SHA-512
69edbedbb687ec163205145dcdde8cb22743546141de43649521d50533f009fb027b737fb80c65a0a88c80c38248eef227eb3c94551983c03b9880c870c62b25
java.security.MessageDigest
3346
not checked
si27.gif
local
1554970138391
0
SHA-512
7b01648f57f24864c7d4d72ec825eb7e86b5bc097f84fcd1ad9567d846ab6f904f4da19914aa5a6dfe5f73dd359dd71756331f676e316a7e1125fe9eac1a40df
java.security.MessageDigest
3341
not checked
si28.gif
local
1554970138392
0
SHA-512
ce225866d8ec942c04e86ccec7f35fb9fbd05a088960f603c4aca3302cd9354c0e7874102c85df4f8cb6fde0875e500a417db0cc9a713d3db23c1149ffe59b1d
java.security.MessageDigest
4019
not checked
si29.gif
local
1554970138393
0
SHA-512
9c412ef438db8c7b044b06ae9b8b4df0c45add46cff2c9bd54190aa3e17d66b0a9ea97986814aafc970e3eaa37875f69e7b428dd0f736576e817f58a77759366
java.security.MessageDigest
2907
not checked
si3.gif
local
1554970138394
0
SHA-512
f5acf3cf5aa5037294c3a49f7d952f38e5a474378f864faba04d639e187c23b0ade411ed2ce3fa8c8041303045826d835133a898ce31091c010c52298f1f9e4b
java.security.MessageDigest
4691
not checked
si30.gif
local
1554970138395
0
SHA-512
f9c498b00a0f3a4238fa55abc11027aca8b7e422afe189d3fd3dbbde81e1b404314adbe17701de7e3069105b88dd6da902bf6a266abff7c7ba447102e7c8ab97
java.security.MessageDigest
2899
not checked
si31.gif
local
1554970138396
0
SHA-512
4d6b7824d80ade57a42ac38e198c7ef240281d20298d1392fda902af325a133f95329c536214a8b3aa43c0940eb0acb884631fce7db3939665b8b786e1416686
java.security.MessageDigest
2931
not checked
si32.gif
local
1554970138397
0
SHA-512
a7a5a92b4555d7516101591ecffcbaa3f62d050ff7920d2d9fbfbe149d098ec495a1b7d9e22c0907e9b2716ba17a70f6b37f62d625a500d079a36801de2cf62e
java.security.MessageDigest
3127
not checked
si33.gif
local
1554970138398
0
SHA-512
73ef07973b40d781397b36ce96c191d42a52c0fb5e6f52b38f37bb9bf5952525de317da9ae1bc1fe9b38e7c17d6672d6e914077582b906845af94a8491b6523b
java.security.MessageDigest
3186
not checked
si34.gif
local
1554970138399
0
SHA-512
f31f17897d00c89e04af99c95ef3cc52f48d115c02b39660fc012ea541c7d300d63cf1c26851a9bd36edd37270a30c9182e6c804b2c2a403b7b8fde81b6588b1
java.security.MessageDigest
3162
not checked
si35.gif
local
1554970138400
0
SHA-512
8e996aa93b14ae1751aba7d1d91055e351f45e431661a97db7b84122c44bc0398f2e85c9f00932ac797527d1d722ba462972aec30ee7e77134e239c47283bd1a
java.security.MessageDigest
3203
not checked
si36.gif
local
1554970138401
0
SHA-512
2624a2c09a9102615403a6e0dcfe1893f875eafa8d30df423de1b15250c6dac836786c71f42050829e83cda255e9f6f8f7d7772b10f76172bd7528e97e3044b0
java.security.MessageDigest
3797
not checked
si37.gif
local
1554970138402
0
SHA-512
39bfebba5a217efe2f3471c15c8302095ecd35a91f1997a2e6443c9604d8da6531d484e37c58931ee32eb7edfa4032f4343a45b94834f5bcd965202fe439bb29
java.security.MessageDigest
4516
not checked
si38.gif
local
1554970138403
0
SHA-512
ef4eac3a18db96f8083cb8de9af6a1b9918c198680bdb64d2e0355d1b936e945441903d62b67657eccc9453d38230b2c5f118004fe8fa3a39d2d932749ec773e
java.security.MessageDigest
2979
not checked
si39.gif
local
1554970138404
0
SHA-512
028cfcdfb19748fc77008d6dade31a346a3b716bbacdb04e74e34f0a8b0e90c37dd5e8ec4f37c49f6c08079ab745e22a1dfe391d114b21e47755ba622b50e236
java.security.MessageDigest
2858
not checked
si4.gif
local
1554970138405
0
SHA-512
d4a4284be4979e7dbade7779ce9cdd1115ce8cce1dfd8a73717f2fc30b32fc6252925a51a3e026bc748e82d95197b33b9b552de7e377ef6a569a78915d361229
java.security.MessageDigest
4448
not checked
si40.gif
local
1554970138406
0
SHA-512
fe7aa0970c2a4ea667b0e9021f7e46fa3be57fbd7ae19611314718a2e9a2fd074af3b76f59d2aff81be3f69d395077870f7871bb9c4916fef0df98920b583f88
java.security.MessageDigest
4467
not checked
si41.gif
local
1554970138407
0
SHA-512
e5f980b7708487869f868dc46754f5356eb138c4217fccbc623edb4f7aebb5ed4071ab25e98c9826e255fb5e3ebfccb35fb2e59f464dfe57e8bd50679ea5de91
java.security.MessageDigest
2848
not checked
si5.gif
local
1554970138408
0
SHA-512
6c436f375e1e0fb130d1d4ead05844d84d29016ec14b73c37eed0e8f6b43632f6074a63d2c90b5225f8b4ae10b1aefe5b51a94dcc7ad9aae3999ab0bd71cf4f5
java.security.MessageDigest
2839
not checked
si6.gif
local
1554970138409
0
SHA-512
2b8b2f89f26ba0d66dbc94c213af88ecf48c230bc81c6620a7cb7eb5c8826fa530188a23a0f7bb276a7f9787da1febc65ae2e38891e94e8e00b00c8ad790f723
java.security.MessageDigest
2841
not checked
si7.gif
local
1554970138410
0
SHA-512
2bb36d170816b0245a014b002dda61fe99bc5efb5e2c0de4be08859cef48ac10c3f629cfebb0e01f8e848533cb846448efe3f5f471f0d2c9c79144fe9e22d663
java.security.MessageDigest
2806
not checked
si8.gif
local
1554970138411
0
SHA-512
62ceba64d7512d7de128a9601b43ac1ab6489fb98a16b9d825ef49f1cd5bd0a0e3f1aeaf9797adbc20a06a1ab9e24232dca9ef3965ef4f3875dca642a79cbdcf
java.security.MessageDigest
2892
not checked
si9.gif
local
1554970138412
0
SHA-512
8afaced5888ba392d0fe4367bbce03de22547bc05609c91b5ebe067dd1e43ddd484fc58f137532073c04531df22de42884a2f110b7acd8f7f84c69570f0fb176
java.security.MessageDigest
41399
not checked
fx1.jpg
local
1554970138413
0
SHA-512
c6cee04b15e7be8acdb2a059747b5e331efc243a216fb77fc220a359d6192ac862991ca1b6815bf11df959d41a72862ca8c2753ff518103ff5ce76b3051c6b02
java.security.MessageDigest
14456
not checked
fx1.sml
local
1554970138414
0
SHA-512
c975cd0fcff517f9266ff13e70f10d28635529c3a45f85cc29975f04ead34abfa6503f30f021f3c30f45b69477999fd89d130d6c8a238e0bb3be19a314a97790
java.security.MessageDigest
82262
not checked
gr1.jpg
local
1554970138415
0
SHA-512
301ddd387fbda4fefe33145eea9ee6dcbb8c7e18e80f36993efb448b0598d7cb6105a5f3e6ecd03c34e24f68ff55b0441c16da586f54ac433100fcb07c452442
java.security.MessageDigest
17709
not checked
gr1.sml
local
1554970138416
0
SHA-512
adca4e972b774ca36e40a151a000cc15a34365d23b66ea889ec2c4f8f42707da28079c56ff7d210cfd90b5a7bb59b23c797f8afac8f9b7998652e9ef85610137
java.security.MessageDigest
99849
not checked
gr2.jpg
local
1554970138417
0
SHA-512
2ddcbdb09ab49a63e084d0e1e1f5a27e24e9f0c097501987086915bdcaa8596f94ca5ffb12dbf649c4967746e9e3998636b39f374cf0d54ea941c8d2709fc4da
java.security.MessageDigest
20615
not checked
gr2.sml
local
1554970138418
0
SHA-512
e19d39afba082699e04049e0aac6c8e798a87b495fc039d377301cb693a2964d9d3f754cf336b728f27649429f5b82ca5ef8562e03300b0b2c41b01360f7c292
java.security.MessageDigest
148694
not checked
gr3.jpg
local
1554970138419
0
SHA-512
b523268fb806f0dedbb73ac4a7d289bfecd10c016747c5b6cadf5b5b5de9385f67437c5965d52c778f4662dacbb85413cb2d75645d0558226fc75d28d59ddb0a
java.security.MessageDigest
25539
not checked
gr3.sml
local
1554970138420
0
SHA-512
57b86b68555f71a28048fb50a1a7569e93166dba5d071765139c62d2a81e585a4f2f89b052ca0b4bc0b1e5d7469677d93e818b2c2d582641d2f95653e16d52cf
java.security.MessageDigest
35453
not checked
gr4.jpg
local
1554970138421
0
SHA-512
4d2ea71be16c85df82f19f71b14f49e4735f20d22a1957419f6065b7829cf0cbfa5a173b2401f317599d7e6f6a09c2612d59341b2fc0c0695b56eb4f64f5ea35
java.security.MessageDigest
14495
not checked
gr4.sml
local
1554970138422
0
SHA-512
6c53d9b9fe8743c8ba87cdba3d72343aedf8c5c29d574cf5ed8f48a4100ecec5fc12bab5401bcf7e866bf475a4b439fc7018fc1d0e52aae8dc8ae05386cb2b4c
java.security.MessageDigest
117946
not checked
gr5.jpg
local
1554970138423
0
SHA-512
a040dc669ead6ca21290aed2dcb40811c3d7853c81b02e593a63d98e1cfa68f3199b3e5e6a3e2acf1c6eda4ed9466ebe8cbaf53899dc4d8a8d62db92e8589090
java.security.MessageDigest
21796
not checked
gr5.sml
local
1554970138424
0
SHA-512
4f170b939f7450818a3f7772cfb314e8b12fccfaea2ceccb463a63d5983c0d7c03dab252dc7438f0ef60f054ffd4f54dd95edfe8997b48fdbba3a9623d49885b
java.security.MessageDigest
1497160
not checked
mmc1.pdf
local
1554970138425
0
SHA-512
e3bf71b1d85d6b4afba98f732ee58951438da6abdd44ed9c3cd75fcfa868bdb05778a123f1bd41fae718ce25bd927ef14448f929dd4dd2b1aeab34c596f06492
java.security.MessageDigest
22443
not checked
metadata.xml
free
00001
The Authors
KB-agent-id
1
supplier
KB-owner-id
00001
KB-agent-id
1
Elsevier
organization
The Authors
ingestion
2019-04-11T09:55:36.754+02:00
Connector
software
Digitaal Magazijn release 1.5
ejournals_esp_1
streamprofile
ingestion2019-04-11T10:07:17.428+01:00Generic IngestsoftwareDigitaal Magazijn release 1.5