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Arsenic mobilization from historically contaminated mining soils in a continuously operated bioreactor: Implications for risk assessment Liwia Rajpert, Boris A. Kolvenbach, Erik M Ammann, Kerstin Hockmann, Maarten Nachtegaal, Elisabeth Eiche, Andreas Schaeffer, Philippe F.-X. Corvini, Aleksandra Sklodowska, and Markus Lenz Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02037 • Publication Date (Web): 25 Jul 2016 Downloaded from http://pubs.acs.org on July 31, 2016
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A
B
C
Figure 1. Mobilization of As (A), Fe (B) and Mn (C) in terms of elemental mobilization rate (circles, primary Y-axis) and total element mobilized (triangles, secondary Y-axis).
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A
B
Figure 2. XRF elemental mapping (1 × 1.5 mm) of the initial (A) and reduced (B) soil with arsenic hotspots (As ≥ 0.3% wt) in overlay (dashed lines).
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normalized absorbance
intial soil
reduced soil
As(V)
As(III) FeAsS
11860 11870 11880 11890 energy [eV]
Figure 3. Normalized As K-edge XANES spectra of initial and reduced soil. Vertical lines mark main edge crest energies of reduced (FeAsS, arsenopyrite) and oxidized (As[III], NaAsO2; As[V], Na2HAsO4 × 7H2O) species.
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Figure 4. Arsenite (open circles), arsenate (solid circles) effluent concentration and number of copies of arrA normalized to 16S rRNA genes (solid squares, secondary Y axis).
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200
0.4
DOC [mg L-1]
0.0 100
-0.2 -0.4
50 -0.6 0 0
500
1000 1500 Time [hours]
2000
Redox potential [V]
0.2 150
-0.8 2500
Figure 5. Dissolved Organic Carbon (DOC) concentrations in reactor effluent (circles, primary Y axis) and normalized redox potential (squares, secondary y axis).
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Table 1. Linear regression analysis using multiple independent variables (top) and Mn as sole independent variable (bottom). Coefficients Standard Error t Stat
P-value
Intercept
1913.1
169.14
11.31
2.53 × 10-11*
Mn [µg L-1]
0.578
0.06
8.94
2.92 × 10-09*
Fe[II] [µg L-1]
1.998
1.41
1.42
0.1683
DOC [µg L-1]
-7.488
3.09
-2.42
0.0231
Redox [mV]
1076.246
493.59
2.18
0.0388
R² = 0.835; adjusted R²= 0.809; ANOVA F = 31.7; p-value: 1.86 × 10-09
Intercept
1621.8
118.09
13.73
5.80 × 10-14*
Mn [µg L-1]
0.499
0.05
9.68
1.96 × 10-10*
R² = 0.770; adjusted R²= 0.762; ANOVA F = 93.7 p-value: 1.96 × 10-10
* highly significant (< 0.001)
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Table 2. Elemental composition of spots containing elevated concentrations of As (≥ 0.3% wt) alone or in conjunction with elevated Mn (≥ 0.14% wt) and/or Fe (≥ 8% wt) concentrations. Initial soil
Reduced soil
Hotspot total area composition
No. of spots
total area No. of spots
[%]
[%]
As
1
0.34
7
1.43
As+Mn
3
0.21
1
0.28
As+Fe
2
0.52
1
0.40
As+Mn+Fe
10
3.34
1
0.56
Sum
16
4.41
10
2.67
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mobilization
from
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1
Arsenic
historically
2
contaminated mining soils in a continuously
3
operated bioreactor: Implications for risk
4
assessment
5
Liwia Rajpert†, Boris A. Kolvenbach†, Erik M. Ammann†, Kerstin Hockmann‡, Maarten
6
Nachtegaal§, Elisabeth Eiche#, Andreas Schäffer¥, Philippe F.-X. Corvini†, Aleksandra
7
Skłodowska˟, Markus Lenz†,ǁ,*
8 9
†
Institute for Ecopreneurship, School of Life Sciences, University of Applied Sciences and
10
Arts Northwestern Switzerland, Gründenstrasse 40, 4132 Muttenz, Switzerland
11
‡
12
Zürich, Universitätstrasse 16, 8092 Zürich, Switzerland, current address: Southern Cross
13
GeoScience, Southern Cross University, 1 Military Road, Lismore 2480, Australia
14
§
Paul Scherrer Institute, 5232 Villigen – PSI, Switzerland
15
#
Institute of Applied Geosciences, Karlsruhe Institute of Technology (KIT), Adenauerring
16
20b, 76131 Karlsruhe, Germany
17
¥
18
Aachen, Germany
19
˟ Laboratory of Environmental Pollution Analysis, University of Warsaw, 02-096 Warsaw,
20
Poland
21
ǁ
22
Wageningen, The Netherlands
Institute of Terrestrial Ecosystems, Department of Environmental Systems Science, ETH
Institute for Environmental Research (Biology V), RWTH Aachen University, 52074
Sub-Department of Environmental Technology, Wageningen University, 6700 EV
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23 24
*
25
TOC ART
corresponding address:
[email protected], (T) +41 614 674 791, (F) +41 614 674 290
26 27 28
Abstract: Concentrations of soil arsenic (As) in the vicinity of the former Złoty Stok gold
29
mine (Lower Silesia, southwest Poland) exceed 1000 µg g-1 in the area, posing an inherent
30
threat to neighbouring bodies of water. This study investigated continuous As mobilization
31
under reducing conditions for more than 3 months. In particular, the capacity of autochthonic
32
microflora that live on natural organic matter as the sole carbon/electron source for mobilizing
33
As was assessed. A bi-phasic mobilization of As was observed. In the first two months, As
34
mobilization was mainly conferred by Mn dissolution despite the prevalence of Fe (0.1% wt
35
vs. 5.4for Mn and Fe, resp.) as indicated by multiple regression analysis. Thereafter, the
36
sudden increase in aqueous As[III] (up to 2400 µg L-1) was attributed to an almost quintupling
37
of the autochthonic dissimilatory As-reducing community (quantitative polymerase chain
38
reaction). The aqueous speciation influenced by microbial activity led to a reduction of solid
39
phase As species (X-ray absorption fine structure spectroscopy) and a change in the elemental
40
composition of As hotspots (micro X-ray fluorescence mapping). The depletion of most
41
natural dissolved organic matter and the fact that an extensive mobilization of As[III]
42
occurred after two months raises concerns about the long term stability of historically As-
43
contaminated sites. 2 ACS Paragon Plus Environment
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Introduction
46
Arsenic (As) is an element with a low crustal average abundance (2–3 µg g-1)1 that is
47
unevenly distributed in the environment. Often, elevated As concentrations are associated
48
with mining activities because As is commonly found in copper, lead, or gold ores.2,3 The
49
Złoty Stok mine (Lower Silesia, southwest Poland) was founded in 1273 and was one of the
50
main gold suppliers for Europe from the 17th century until its closure in 1962.4–6 The basic
51
rock in the mine is composed of mica schists, mica-quartz schists, and quartzite schists7 while
52
the main As-bearing minerals are loellingite (FeAs2), scorodite (FeAsO4 × 2H2O), and
53
arsenopyrite (FeAsS).8 These mining activities generated a large amount of As-rich waste
54
material, which was disposed of in nearby valleys, including the Trująca Valley.5–7
55
Occasional seepage and dam overflow affected parts of the valley localized downstream from
56
the tailings disposal areas,9,4 which caused As soil contamination exceeding 1000 µg As g-1.5,4
57
Mobilization of As from its solid host phases into surface and groundwater caused increased
58
As concentrations ranging from 990 µg As L-1 (springs located on the deposit’s periphery) to
59
26160 µg As L-1 (water-draining audits).6 Poisoning from As in this region was reported as
60
early as the 19th century and described as “Reihensteiner Krankheit” (Złoty Stok Disease).2,10
61
Mobilization of As causing contamination of natural waters is understood to occur
62
primarily through four mechanisms: (I) via direct desorption under alkaline conditions; (ii) via
63
reductive dissolution of As bearing minerals; (iii) via oxidation of reduced As-S-minerals and
64
(iv) via geothermal waters. 11 Microbes may play a catalytic role in mobilization, for instance
65
through dissimilatory reduction of sorbed As[V] and subsequent release of As[III] in the
66
aqueous phase.12
67
For the Złoty Stok area, the main source of As-contaminated groundwater stems from an
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oxidative, microbe-catalysed dissolution of As-bearing minerals.13–15 Mobilization of As that
69
is driven by the reductive dissolution of Fe phases has received a great deal of attention in the 3 ACS Paragon Plus Environment
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context of the mass poisoning in southeast Asia,12 and was presumed to be dominant in
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different lake and river sediments,16–18 shallow and intermediate aquifers,19 and paddy
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soils.20,21 In contrast to that, the reductive dissolution of Mn phases has been studied to a
73
lesser extent.22–26 Whereas the microbial diversity of mine biofilms and their impact on As
74
mineral dissolution has been studied well,15,27,28 mobilization of As by autochthonic soil
75
microbial communities has not been extensively investigated. In particular, studies are scarce
76
on As mobilization in soils impacted by old mines under reducing conditions. Such reducing
77
conditions may be induced locally through changes in soil use (compaction and resulting
78
water-logging) and at larger scale throughflooding events29. In addition, heterogenic soil
79
aggregates within the bulk soil may bear redox / chemical gradients, allowing for spatially
80
confined anoxic biogeochemical reactions30
81
Therefore, the present study investigated the mechanisms that govern As mobilization
82
from heavily contaminated soil in a continuously operated bioreactor over a period of more
83
than three months. A continuously operated bioreactor was used to distinguish the
84
contribution of solid phase reductive dissolution (Fe, Mn phases) and dissimilatory As
85
reduction on As mobilization in time, which is hard to accomplish in batch experiments. A
86
combination of speciation methods in the solid and liquid phases, i.e., liquid chromatography-
87
inductively coupled plasma mass spectrometry (LC-ICP-MS), X-ray absorption near edge
88
structure spectroscopy (XANES), elemental soil mapping (micro X-ray fluorescence, µXRF),
89
and molecular biology (quantitative polymerase chain reaction, qPCR) was applied. Emphasis
90
was given to elucidating the role on As mobilization of the autochthonic soil microflora that
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feed only on natural organic matter.
92 93
Materials and Methods
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Soil sampling
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The soil originated from the gold mine in Złoty Stok (Lower Silesia, southwest
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Poland). Samples (~20 kg) were collected from the surface layer (0-20 cm) from one point on
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the bank of the Trująca River (“Poisonous Stream”), into which the former gold mine
98
drained.4,6 The soil was air-dried, sieved (mesh size 2 mm), and homogenized using the cone
99
and quarter technique.31 This soil is hereinafter referred to as “initial soil”, whereas the soil
100
obtained after a reactor operation is referred to as “reduced soil”.
101 102
Bioreactor operation
103
Mesophilic (21 ± 5 °C) completely stirred 1 L tank reactor (Multifors, Infors HT,
104
Bottmingen, Switzerland) was inoculated with soil (10% w/v) and operated for about 100
105
days at a hydraulic retention time of 48 h at a slightly basic pH (pH = 8.0 +/- 0.6, using
106
automated acid/base dosing). The medium contained no external organic carbon source (SI).
107
Anaerobic conditions were maintained by providing a continuous N2 flow in the reactor
108
headspace. The off gas was directed through two gas traps (1% H2O2,32 and 69% HNO333) to
109
capture potentially formed As-volatile species. Fresh medium was stored at 4 °C, filtrated
110
(0.22 µm filters, Merck Millipore, Molsheim, France) and continuously supplied to the reactor
111
by a peristaltic pump. The reactor effluent was pumped through a self-made, low-density
112
polyethylene filter to a gravitational settler (additional hydraulic retention time of 72 h) in
113
order to separate fine soil particles. Subsequently, the liquid was transported from the settler
114
to a waste container.
115 116
Liquid phase analysis
117
Samples (~20 mL) were collected from the reactor, centrifuged (4500 rcf, 10 min, 21
118
°C), and the resulting supernatant was syringe filtered (0.45 µm and 0.2 µm pore size,
119
Whatman, Hertogenbosch, The Netherlands). The totally dissolved element concentrations
120
were analysed on an Agilent 7500cx ICP-MS (SI).Samples for As speciation were preserved 5 ACS Paragon Plus Environment
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with EDTA/AcOH34, stored at 4 °C, and measured within 72 h after sampling. For detection
122
of As species, i.e., arsenite (As[III]), arsenate As[V], monomethylarsonic acid (MMA[V]),
123
and dimethylarsinic acid (DMA[V]), a LC system (SI) was coupled to the ICP-MS. Dissolved
124
iron (Fe[II]) was determined by a spectrophotometric method with 1.10-phenantroline35 using
125
0.5 M HCl to preserve the original speciation. Dissolved organic carbon (DOC) was detected
126
with a Liquid Chromatography-Organic Carbon Detection Model 8 (LC-OCD, DOC Labor
127
Huber, Karlsruhe, Germany)36
128 129
Solid phase characterization
130
Total element soil concentrations were determined using an energy dispersive X-ray
131
fluorescence (EDXRF) analysis (Ametek SpectroXRF Xepos III, Spectro, Kleve, Germany).
132
Major crystalline soil phases were characterized by X-ray powder diffraction (XRD) on
133
milled samples using a Bruker AXS D8 (Bruker AXS GmbH, Karlsruhe, Germany) Advance
134
system (CuKα1) equipped with a Lynxeye superspeed detector at a scanning speed of 0.009°
135
s-1. Elemental mapping in the energy range 2–20 keV was performed on a bench scale µ-XRF
136
instrument (µ-Eagle III, Röntgenanalytik GmbH, Taunusstein, Germany) equipped with a
137
Si(Li) detector. Excitation conditions in the rhodium (Rh) X-ray tube were set to 50 kV and
138
500 mA. An additional 50 µm-thick Rh filter was set into the path of the X-rays to eliminate
139
most of the Bremsstrahlung and to provide nearly monochromatic X-rays with the
140
characteristic RhKα and RhKβ radiation. The beam was focused by a polycapillary lens to a
141
spot size of about 30 µm in diameter. Areas of 1 × 1.5 mm2 were scanned with a 30 µm step
142
size with 100 s (initial) and 50 s (reduced) dwell time. PyMca software37 was used for the
143
analysis of the received X-ray intensity-energy spectra. Prior to the analysis, the soil samples
144
were pellet pressed (Manual Hydraulic Press 15–25 Ton, Specac Limited, River House,
145
England) and moulded with EpoMet®G hot mounting epoxy compound (SimpliMet 2
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Mounting Press, Buehler, Lake Bluff, USA). Subsequently, polished sections of each sample
147
were prepared using an optical polisher (Tegra-Pol 21, Struers A/S, Ballerup, Denmark).
148
The arsenic K-edge spectra of soil samples were recorded at the SuperXAS beamline
149
of the Swiss Light Source (SLS, Paul Scherrer Institut, Villigen, Switzerland). The
150
measurements and data processing is detailed in the supplementary information.
151 152
Metagenomic DNA isolation
153
Soil samples were collected into 50 mL falcon tubes (Greiner Bio-One GmbH,
154
Frickenhausen, Germany) that had been rinsed beforehand with a 10% (v/v) NaClO solution
155
in sterile, nanopure water to remove any DNA.38 Soil slurry aliquots were centrifuged (4500 ×
156
g, 15 min), the supernatants were discarded, and the pellets were stored at -20 °C until
157
analysis. Metagenomic DNA was isolated from the soil pellets according to Miller et al. 38 and
158
Sagova-Mareckova et al.,39 with minor modifications (SI). The purity of the extracted DNA
159
was measured using a UV-VIS spectrophotometer (SI, Table S1) (NanoDrop ND-1000
160
Spectrophotometer, Thermo Scientific, Wilmington, USA) and analysed by means of
161
electrophoresis (BioRad Sub-Cell®GT, BioRad Laboratories, Hercules, USA).
162 163
Quantification of the 16S rRNA genes and dissimilatory arsenate reduction (arrA) genes
164
Quantification of 16S rRNA (~180bp) genes was performed with the primers
165
described by Clifford et al.40 Each 25-µL reaction volume contained 12.5 µL of SYBR Green
166
qPCR Master Mix, 40 nM of each primer, 1 µL of DNA templates of known concentrations of
167
standards (10-1–10-8) or 1 µL of 100 × diluted DNA extracted from soil samples. The PCR
168
protocol was 15 min at 95 °C, with 45 cycles consisting of 10 s at 95 °C, 20 s at 58 °C, and 1
169
min at 72 °C, then 50 s at 72 °C. The amplified 16S rRNA gene (1466bp) from Chrysiogenes
170
arsenatis BAL-1 in dilutions of 10-1–10-8 was used for preparation of the standard curve.
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Quantification of arrA (~180bp) genes was performed with primers described by
172
Malasarn et al.41 Each 20 µL reaction contained 10 µL SYBR Green qPCR Master Mix, 1.25
173
mM MgCl2, 0.5 µM of each primer, and 1 µl of DNA templates of known concentrations of
174
standards (10-1–10-8) or 1 µl of 100 × diluted DNA extracted from soil samples. The
175
amplification cycles included an initial denaturation at 95 °C for 15 min followed by 40
176
cycles of denaturation at 95 °C for 30 s, annealing at 52 °C for 40 s, and extension at 72 °C
177
for 1 min. The amplified arrA gene (893bp) from Chrysiogenes arsenatis BAL-1 in dilutions
178
of 10-1–10-10 was used for preparation of the standard curve (SI).
179 180
Statistical Data Treatment
181
A multiple linear regression analysis was performed to determine possible correlations
182
between total As, Fe[II], Mn, and DOC-effluent concentrations as well as the redox values
183
using Microsoft Excel (Microsoft Corporation, Redmond, USA). These were done over the
184
entire reactor operation (SI, Figure S5, Table S2) and during the first 1400 h only (Table 1, SI,
185
Figures S6-7). Local polynomial regression fitting (LOESS) was employed on all replicate
186
measurements to smooth the As mobilization rate (SI, Figure S1).
187 188
Results
189
Continuous element mobilization
190
Mobilization of As followed two distinctive phases: 0–1400 h (phase I) and 1400–
191
2352 h (phase II). Their identification was based on the total As mobilization rate (Figure 1A,
192
SI, Figure S1), the increase in As[III] and arrA gene copies (Figure 4), redox potential, and
193
DOC concentration (Figure 5).
194
Phase I was characterized by an initially high As mobilization rate (max. 94 µg L-1 h-1
195
at 72 h) and a gradual decrease until ~1400 h, with a minimal mobilization rate of 25 µg L-1 h-
196
1
(912 h, Figure 1A). In this timespan, 685 µg g-1 of As (~35% of total) were mobilized from 8 ACS Paragon Plus Environment
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the soil. Phase II started at ~1400 h, with a maximal mobilization rate of 83 µg L-1 h-1. After
198
~1600 h, the As mobilization decreased considerably and reached a minimum mobilization
199
rate of 4 µg L-1 h-1 after 2088 h. (Figure 1A). Overall, 990 µg g-1 (corresponding to 49% of the
200
total As) were mobilized during reactor operation.
201
Initially, Fe was mobilized to an average of 28 µg L-1 h-1 (0-576 h). At ~700 h,
202
mobilization peaked and 120 µg L-1 h-1 were released for a short period of time (Figure 1B;
203
SI, Figures S2-3). Afterwards, Fe was mobilized at a considerable lower rate (0.16% Mn, Figure 2
248
middle).
249
The As K-edge spectra of the initial and reduced soil were clearly different in the
250
XANES region. In the initial soil dominated features at the energy characteristic for more
251
oxidized species (11873.8 eV for As[V]). Whereas in the sample taken upon termination of
252
the reactor operation appeared a distinct feature at the energy characteristic for more reduced
253
species (11870.0 eV for As[III]) (Figure 3). Accordingly, the best LCF that was obtained in
254
the initial sample yielded scorodite almost exclusively (11874.6 eV, 96%, SI, Figure S10),
255
which matches the mineralogy of the Złoty Stok region.8 In the reduced soil, the best LCF that
256
was obtained showed a ~60% contribution of the reduced species (the best fit yielded
257
arsenolite and As[Cys]3) and only a ~40% contribution of scorodite (SI, Figure S10, Table
258
S3). However, despite the large amount of possible reference compounds used, some residual
259
components could not be fitted. As a result, species structurally similar to scorodite may
260
contribute to the soil speciation as well, particularly because As K-edge XANES features can
261
be very similar for different model compounds43,44.
262 263
Development of a dissimilatory arsenate-reducing bacterial community
264
Under reducing conditions (0.0–0.4 V) occurring within the reactor, arsenite was the
265
main species detected (Figures 4-5). Mobilization of both arsenite and arsenate was initially
266
high (effluent concentrations up to ~3100 µg L-1 As[III] and ~2400 µg L-1 As[V]), and
267
decreased during phase I. After ~1000 h, the increase in the As mobilization rate (Figure 1A)
268
was mainly related to arsenite (max. As[III] concentration 2398 µg L-1 at ~1500 h). Within the
269
same time frame, an increase of arrA gene copies was detected. The latter is a reliable marker
270
for metabolically active Dissimilatory Arsenate Reducing Bacteria (DARBs).41 The highest
271
amount of arrA gene copies that had normalized to an overall number of 16S rRNA gene
272
copies was observed at 1536 h (max. arrA gene copy number 1.57-3). Subsequently, the 11 ACS Paragon Plus Environment
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considerable decline of arrA gene copies number was observed (min. arrA gene copy number
274
3.00-4 at1872 h, Figure 4). The increase of DARB activity was accompanied by a peak in
275
DOC concentration and an increase in redox potential (~1500 h, Figure 5).
276 277
Discussion
278
Continuous As mobilization from historically contaminated soil
279
Under reducing conditions, the indirect arsenate mobilization via dissolution of Fe[III]
280
(oxyhydroxides) and oxides is considered as one of the main routes of As release in sediments
281
and water-logged soils.12,45 This study gives evidence that, despite their considerably lower
282
soil concentrations, the reductive dissolution of Mn phases was the main factor controlling As
283
release for more than 2 months (phase I). When both Fe and Mn pools that were prone to
284
reductive dissolution were depleted, the mobilization of As was ultimately controlled by the
285
activity of autochthonous DARBs that were developing on natural organic matter.
286
Surprisingly, this occurred suddenly and after a 2-month reactor operation under constant
287
conditions. Due to historical As soil contamination, potential autochthonic DARBs were
288
believed to have developed over long periods of time2,4,6 and available As directly depleted.
289
The dynamics of As continuous mobilization may be related to a complex interplay between
290
solid As-bearing phases and bio-induced changes in speciation. In the initial soil, the elements
291
As, Fe, and Mn co-occurred as indicated by µXRF mappings (Figure 2AB, Table 2), although
292
no crystalline As/Fe/Mn mineral phases were detectable using XRD (SI, Figure S9). The
293
XANES analysis demonstrated a high (96%) contribution of oxidised As species in the initial
294
soil (Figure 3). The best LCF yielded mainly scorodite (FeAsO4 × 2H2O), which is one of the
295
most common natural arsenates in acidic, oxidizing environments (SI, Table S3). Furthermore
296
the presence of scorodite as an oxidation product of arsenopyrite or loellingite46–48 is in good
297
accordance with previous studies on the mineralogy of the region.4–6. The primary dissolution
298
phase was characterized by a high As mobilization rate (Figure 1A), which was accompanied 12 ACS Paragon Plus Environment
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by a rapid Fe mobilization within the first ~800 h (Figure 1B). However, multiple linear
300
regression analysis yielded dissolved Mn alone as a significant factor (p = 2.9×10-9)
301
influencing As concentrations until ~1400 hours of reactor run (Table 1; SI, Figure S7). In
302
contrast to other studies,18,49,50 neither dissolved Fe (SI, Figures S2-3) nor DOC (Figure 5)
303
impacted As mobilization from the soil (p = 0.17 and p = 0.02, respectively) (Table 1). The
304
fact that Mn mobilization steadily increased during the first ~3 days after the start suggests
305
that the dissolution is bio-mediated. Due to the similarity in As K-edge XANES features of
306
different model compounds43,44 species structurally similar to scorodite may well contribute to
307
the soil As speciation. Thus, we rather assign As being associated to Mn and/or mixed Fe-Mn
308
phases than pure scorodite, which is supported by µXRF mappings as well (Table 2).
309
Manganese-oxides have been reported to affect overall As mobility via adsorption and
310
oxidation of arsenite to arsenate,45 which may explain the occurrence of arsenate (~24% of
311
total released arsenic) in the liquid phase despite the reducing conditions (Figure 5) and
312
DARB activity (phase II, Figure 4). Those results are in good agreement with Lafferty et al.,51
313
who showed that As, once adsorbed to δ-MnO2, exists only in the form of arsenate. Though
314
the precise mechanism remains to be discovered, the use of a continuous bioreactor in contrast
315
to batch studies had the clear advantage of separating the underlying processes in time. In
316
batch experiments, steadily increasing element concentrations may lead to a misinterpretation
317
of the dominant processes.
318
During phase I, low redox potentials (-0.2–-0.4 V) and high concentrations of DOC
319
(Figure 5) constituted a readily available carbon source for developing autochthonic
320
microflora. In consequence, a sudden increase in As mobilization was observed after more
321
than 2 months of operation. Since As mobilization was mainly related to As[III] (Figure 4)
322
and an almost fivefold relative increase in the number of arrA gene copies was detected
323
(Figure 4), this could be attributed to the activity of indigenous DARBs. The higher redox
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324
potential measured during this period was consistent with the As[V]/As[III] couple as the
325
dominant redox reaction.52,53
326
Upon the end of the reactor operation (~1700 h), a small amount of As was mobilized,
327
which could be explained by a depletion of As-bearing Fe/Mn pools and/or formation of
328
secondary As-bearing phases. For instance, the arsenate produced in reaction with Mn-oxides
329
might be subsequently adsorbed to Fe[III]/Mn-(oxyhydr)oxides and/or precipitated with
330
microbially produced Fe[II] as As-sequestering ferrihydrite.45,51,54–56 Reductive dissolution of
331
scorodite was recently found to yield Fe-As phases (biogenic ferrous arsenite and
332
parasymplesite (Fe3[AsO4]2 × 8H2O).57 However, the µXRF mapping did not indicate an
333
increase of Fe/As rich hotspots in the soil after the reactor was run (Figure 2B, Table 2).
334
Whereas elemental maps showed mainly mixed Mn/Fe/As phases in the initial soil, the
335
reduced soil contained mostly hotspots consisting only of As (Table 2). XANES of the
336
reduced soil demonstrated a more reduced speciation of As in the solid phase (Figure 3),
337
though the exact identity remains unknown due to the inability of the XANES analysis to
338
clearly distinguish between structurally similar As phases.44, 43 Yamaguchi et al.
339
that the share of arsenite in the paddy soil solid phase after anaerobic incubation increased
340
considerably, while the proportion of arsenite in the solid phase in gamma-irradiated samples
341
did not change. This suggests that the presence of As[III] in the soil solid phase of this study
342
may have a biotic origin as well (Figure 3). The autochthonic microflora was either able to
343
reduce arsenate directly in the soil solid phase,21,58 or As[III] released by DARB was
344
secondarily immobilized. The formation of As[III]-NOM complexes might indeed explain
345
why no µXRF amendable element (Fe, Mn, S, Ca, etc.) was found to be associated with the
346
As hotspots in the reduced soil (Table 2, Figure 2).
20
reported
347 348
Implications for risk assessment of historically contaminated soils
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349
Contamination of the Złoty Stok region with As has been under investigation since the
350
1960s.59,60 The emerging role of As-hypertolerant, oxidizing/reducing bacteria that contribute
351
substantially to the release of As from minerals inside the gold mine was recently
352
described.15,27,28 This study adds to the identification of historically contaminated soil as a
353
latent and considerable source of As release. The fact that monitoring total element
354
concentration is an insufficient way to assess contaminant risk has become generally accepted
355
in recent years.61 Therefore, different simplified approaches such as sequential extraction
356
schemes, eluate tests, or in vitro bioaccessibility tests are applied to estimate site-specific
357
risks.62–66 This study questions the applicability of the aforementioned: (i) a substantial and
358
sudden mobilization of As occurred after times much longer than any accelerated test (~60
359
days); (ii) the mobilization was not related to the reductive dissolution of any major As-
360
bearing phase and is thus not accessible to chemical extraction targeting the latter; (iii) the
361
mobilization occurred due to the DARBs that inevitably developed on a refractory pool of As
362
(i.e., after easily accessible As was already leached). Furthermore, the DARB activity resulted
363
in the release of As[III] yet not in formation ofvolatile arsines (LOD = 1 µg g-1 of gas traps),
364
being of high toxicity in comparison to the oxyanions.
365
contaminated soils. 69–72
67,68
The latter are often emitted from
366
The fact that mining activities ceased more than 50 years ago would suggest that most
367
of the potentially easily available As would have already been leached. This study proves the
368
contrary, which is that an induction of reducing conditions, e.g., by changes in soil use
369
(compaction and resulting water logging; heavy irrigation), changes in hydrology and / or
370
seasonal flooding events, certainly poses a threat to potable water contamination of this
371
region. The As deposits in the area of the Złoty Stok gold mine are the largest in Poland and
372
are the main source of surface and groundwater contamination in the Klodzko Valley.73 In
373
spite of this, the lower part of the Trująca Valley is still in agricultural use as arable fields,
374
meadows, and pastures, which increases the risk of As exposure and the possible entry of As 15 ACS Paragon Plus Environment
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375
into the human food chain.4,5,74 Due to an increasing demand for potable water in the Złoty
376
Stok area, use of the surface water (i.e., the Gold Stream) has been considered.6 This study
377
underlines that this requires not only suitable water treatment, but also long-term monitoring
378
programs for historically contaminated soil as one of the main reservoirs of this carcinogenic
379
element.6
380 381
Associated Content
382
Supporting Information
383
Bulk X-ray fluorescence analysis, X-ray diffraction patterns, XANES reference compounds,
384
detailed information on extraction schemes, total element mobilization, molecular biology
385
techniques and statistical analysis can be found in the supporting information. This
386
information is available free of charge via the Internet at http://pubs.acs.org.
387 388
Author Information
389
Corresponding Author
390
*Phone/fax: +41 614 674 791/ +41 614 674 290; e-mail:
[email protected].
391
Notes
392
The authors declare no competing financial interest.
393 394
Acknowledgments
395
The authors thank Christoph Giger for his contribution to the bioreactor design. We thank the
396
participants of the " Cook and Look " course at PSI, during which the XAFS data were
397
recorded and Professor Yungmin Pan (University of Saskatchewan), who kindly provided
398
additional reference spectra. The support of the Rectors' Conference of the Swiss Universities
399
CRUS within the Sciex-NMSch scholarship (11.130) is gratefully acknowledged. Authors
400
thank Utz Kramar for his help during µXRF measurements at KIT. 16 ACS Paragon Plus Environment
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401 402
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