Diffusive Gradients in Thin Films Reveals Differences in Antimony and

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Diffusive gradients in thin films (DGT) reveals antimony and arsenic mobility differs in a contaminated wetland sediment during an oxic-anoxic transition Maja Arsic, Peter R Teasdale, David Thomas Welsh, Scott G Johnston, Edward D. Burton, Kerstin Hockmann, and William Walpole Bennett Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b03882 • Publication Date (Web): 05 Jan 2018 Downloaded from http://pubs.acs.org on January 5, 2018

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Diffusive gradients in thin films (DGT) reveals

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antimony and arsenic mobility differs in a contaminated

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wetland sediment during an oxic-anoxic transition

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Maja Arsic,a Peter R. Teasdale,b,c David T. Welsh,a Scott G. Johnston,d Edward D. Burton,d Kerstin

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Hockmann,d and William W. Bennetta,d*

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a

Environmental Futures Research Institute, Griffith University, Gold Coast campus, QLD 4215,

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Australia b

Natural and Built Environments Research Centre, School of Natural and Built Environments,

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University of South Australia, South Australia 5095, Australia. c

Future Industries Institute, University of South Australia, South Australia 5095, Australia. d

Southern Cross Geoscience, Southern Cross University, Lismore, NSW 2480, Australia

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*Corresponding Author: [email protected]

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Abstract

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Antimony (Sb) and arsenic (As) are priority environmental contaminants that often co-occur at mining-impacted

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sites. Despite their chemical similarities, Sb mobility in waterlogged sediments is poorly understood in

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comparison to As, particularly across the sediment-water interface (SWI) where changes can occur at the

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millimetre scale. Combined diffusive gradients in thin films (DGT) and diffusive equilibration in thin films

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(DET) techniques provided a high resolution, in situ comparison between Sb, As and iron (Fe) speciation and

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mobility across the SWI in contaminated freshwater wetland sediment mesocosms under an oxic-anoxic-oxic

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transition. The shift to anoxic conditions released Fe(II), As(III) and As(V) from the sediment to the water

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column, consistent with As release being coupled to the reductive dissolution of iron(III) (hydr)oxides.

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Conversely, Sb(III) and Sb(V) effluxed to the water column under oxic conditions and fluxed into the sediment

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under anoxic conditions. Porewater DGT-DET depth profiles showed apparent decoupling between Fe(II) and Sb

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release, as Sb was primarily mobilized across the SWI under oxic conditions. Solid-phase X-ray absorption

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spectroscopy (XAS) revealed the presence of an Sb(III)-S phase in the sediment that increased in proportion with

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depth and the transition from oxic to anoxic conditions. The results of this study showed that Sb mobilization was

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decoupled from the Fe cycle and was, therefore, more likely linked to sulfur and/or organic carbon (e.g. most

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likely authigenic antimony sulfide formation or Sb(III) complexation by reduced organic sulfur functional

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groups).

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Keywords: antimony, arsenic, mobility, DGT, DET, sediment, XAS

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Introduction

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Antimony (Sb) and arsenic (As) are toxic, nitrogen-group elements with similar valence electron configurations

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(ns2np3) and chemical speciation. In natural waters, the reduced inorganic species (Sb(III) and As(III)) typically

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exist as the neutral oxyanions Sb(OH)3 and As(OH)3, while the oxidized inorganic species (Sb(V) and As(V)) are

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present as the Sb(OH)6- and H2AsO4-/HAsO42- oxyanions, respectively.1 While the biogeochemistry of arsenic has

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been extensively researched2, 3 due to its global distribution and high toxicity, antimony is comparatively poorly

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studied, despite its emergence as an important contaminant due to increased mining and industrial use.4 The

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mobilization and sequestration mechanisms of antimony in aquatic systems remain poorly understood.5

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Aquatic sediments play an important role as sinks for metals and metalloids.6, 7 To date, the majority of research

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on antimony biogeochemistry has been conducted on oxic environmental systems,5, 8-11 with relatively few studies

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on its behavior in the anoxic environment, such as aquatic sediments or wetland soils. 12-15.16 This has resulted in a

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focus on the role of iron and manganese hydr(oxide) solubility controlling antimony mobility.11 While this is

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relevant to oxic systems, there is a general lack of information how the generation of reduced species, such as

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ferrous iron or sulfide, affects antimony mobility. Recent work by Fawcett and co-workers16 revealed the

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presence of an authigenic antimony sulfide phase in the sediments of aquatic systems adjacent to a former mine

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site, as well as the association of antimony with particulate natural organic matter. A recent study by Bennett et

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al.,17 using extended X-ray absorption fine structure (EXAFS) spectroscopy, demonstrated the importance of

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antimony sorption to sulfur in a sediment sample collected from the same site as that used in this study. Reduced

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organic sulfur functional groups (e.g. thiols) were recently shown to be important in controlling arsenic mobility

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in waterlogged peat sediments,18 and could be similarly important for antimony.

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Sediments are complex, heterogeneous environments where biogeochemical zones can vary at the millimeter

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scale, especially across the sediment-water interface (SWI).19 For example, the penetration of oxygen in organic-

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rich wetland sediments is typically only a few millimeters, due to intense microbial respiration of organic matter.7

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Therefore, to accurately investigate the in situ mobility of antimony in freshwater wetland sediments across this

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fine-scale zone, the sampling techniques employed must create minimal disturbance to redox-sensitive analytes

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and must be able to capture changes in porewater solute distributions at high spatial resolution. Conventional

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sampling techniques, such as core slicing and centrifugation to recover porewater, cannot capture a sufficiently ACS Paragon Plus Environment

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fine resolution to accurately interpret biogeochemical processes in productive wetland sediments.20 Furthermore,

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these approaches can result in unrepresentative data due to porewater mixing and associated chemical reactions

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that occur during sample processing.19, 21, 22 In contrast, approaches such as the diffusive gradients in thin films

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(DGT) and diffusive equilibration in thin films (DET) techniques can provide millimeter-scale, in situ

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measurements of multiple porewater solutes.23

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DGT techniques have been developed for a range of metals and metalloids,24 and have recently been modified to

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allow the measurement of arsenic and antimony speciation.25,

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selectively measuring ferrous iron at 1 mm resolution,27 and can be combined with DGT to provide coincident

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profiles of iron and metal(loid) contaminants.28 Combined DGT-DET probes were previously used in a controlled

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mesocosm study by Bennett et al.

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anoxic shifts in freshwater and estuarine sediments at high resolution across the SWI. As antimony mobility has

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been linked to iron hydr(oxide) solubility in oxic soils, this study will expand on this mesocosm design to

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investigate whether similar mechanisms control antimony behavior in a contaminated waterlogged sediment. This

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study will also compare how oxic-anoxic shifts affect aqueous antimony and arsenic speciation and whether this

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has implications for antimony mobility across the SWI.

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This study compared the mobility and speciation of antimony to arsenic and iron in a contaminated wetland

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sediment using mesocosms that underwent a transition from oxic to anoxic conditions. DGT and DET

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measurements captured changes in dissolved antimony, arsenic, and iron concentrations and speciation at the

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millimeter scale, which is critical for interpreting the fine-scale redox-driven processes that occur in productive

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sediments. The aqueous phase DGT-DET data was complemented by antimony solid-phase speciation using X-

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ray absorption spectroscopy (XAS). The combined DGT-DET and XAS approach facilitated the acquisition of

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highly representative pore-water and solid-phase speciation information, thus enabling a comparison of the

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complex aqueous antimony and arsenic redox changes in the sediment mesocosm study. Further, this technique

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allowed for investigation of the SWI, the most critical zone across which contaminants are mobilized in aquatic

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systems.

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DET techniques have been developed for

to demonstrate the coupled mobility of iron and arsenic species over oxic-

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Experimental

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Sediment collection and mesocosm preparation. Contaminated sediment was collected from a wetland adjacent

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to a former antimony (stibnite) processing plant located in Urunga, New South Wales, Australia (30°30'13.8"S

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153°00'46.0"E) (see Warnken et al. 201729 for a detailed site description). The top 10-15 cm of sediment was

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transferred into two 10 L acid-rinsed plastic buckets and transported to the laboratory where it was sieved to < 2

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mm and thoroughly homogenized. The sampled sediment contained approximately 15,600 mg kg-1 of iron, 150

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mg kg-1 of arsenic and 270 mg kg-1 of antimony, as measured in the dry sediment by aqua-regia microwave

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extraction (CEM Mars 6) followed by ICP-MS analysis (Agilent 7900) as per USEPA Method 3051A.30 Total

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organic carbon concentrations in the wetland sediments averaged 25 ± 7 %.29 Sediments collected from the same

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site, but at a different time, contained 0.7% total sulfur; 82% of which was present as organic sulfur (cysteine,

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sulfonate, sulfoxide) and 18% as inorganic sulfur (elemental sulfur and sulfate) (see Table S7 of Bennett et al.

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201717). Six 9.4 L acrylic mesocosms (30 cm high by 20 cm diameter), each containing approximately 3 L of

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sediment and 6.4 L of synthetic freshwater, were prepared with similar solute concentrations to those reported

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previously for the site (Supporting Information, Table S1).31, 32 Mesocosms were prepared and allowed to stabilize

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for three weeks in the dark, in a constant temperature room (21 ± 1 °C), with the overlying water layer being

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sparged with air to ensure oxygen saturation and sufficient mixing, prior to the incubation.33 Following sediment

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stabilization, three mesocosms were assigned as controls, which remained oxic during course of the experiment

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due to sparging of the overlying water with air. The other three mesocosms, assigned to the oxic-anoxic-oxic

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treatment, defined the progress of the experimental stages: oxic (days 1 – 5), anoxic (days 6 – 20) and oxic (days

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21 to 23). An acrylic lid and waterproof sealant (All Clear, Selleys) was used for sealing on day 5. This allowed

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the sediment oxygen demand to induce a transition from oxic to anoxic conditions in the overlying water of the

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treatment mesocosms. On day 20, the sealant and lid were removed and mesocosm water columns were sparged

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with oxygen once more to induce an anoxic-oxic transition.

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Water column analyses. During the anoxic phase of the treatment, mesocosms were positioned around a small

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electric motor that drove suspended magnetic stir bars in the water column of each mesocosm, preventing the

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formation of a diffusive boundary layer at the sediment-water interface. Daily in situ measurements of dissolved ACS Paragon Plus Environment

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oxygen (Opti-Ox, Mettler Toledo), Eh (LE510 ORP electrode, Mettler Toledo), pH and temperature (LE438 pH

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electrode, Mettler Toledo) in the water column were taken via a sampling port in each mesocosm lid, which was

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sealed with an air- and water-tight rubber stopper during the incubations. Water samples were collected daily

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through the sampling port, filtered through a 0.45 µm pore size syringe filter (Mixed Cellulose Ester; Millipore)

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and preserved with ultrapure HNO3 (Baseline; Seastar) to 2% (v/v). These samples were measured for total

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metals and sulfur by ICP-MS. Dissolved Fe(II) samples (3 mL) were immediately fixed with ferrozine reagent

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prior to the absorbance being measured at 562 nm within 3 hours of collection.34, 35 Dissolved organic carbon

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(DOC) was measured by calibrating the absorbance at 254 nm to the site-specific organic carbon. A high

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concentration sample of DOC from the sediment pore-water was extracted and the concentration measured by

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high temperature combustion, non-dispersive infrared detection. This stock solution was then diluted to create a

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calibration curve with known DOC concentrations and the absorbance at 254 nm of standards and samples

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measured using a UV-Vis spectrophotometer (Shimadzu UV-1800).36 Arsenic speciation was determined using

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anion-exchange solid-phase extraction cartridges, as described previously;37 the eluted As(III) concentrations

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were determined by ICP-MS. Samples for antimony speciation were acidified to pH 5 with ultrapure HNO3

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before adding 1.0% ammonium pyrrolidine dithiocarbamate (APDC) to complex Sb(III).38 The sample was then

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passed through a C8 solid-phase extraction cartridge (SiliaPrep C8; Silicycle); the eluted Sb(V) concentrations

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were determined by ICP-MS. As(V) and Sb(III) concentrations were calculated as the difference between the

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total concentrations in grab water samples and the concentrations in the solid phase extraction eluents (see Table

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S2 for further details).

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Preparation of DGT-DET samplers. Sediment DGT probe housings were purchased from DGT Research Ltd.

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Mercapto-silica binding gels for the selective measurement of As(III) and Sb(III) were prepared as described by

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Bennett and co-workers.25, 26 Metsorb-Chelex mixed binding layer (MBL) binding gels for the measurement of

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total inorganic As and Sb were prepared as described by Panther and co-workers.39 The diffusive layer of the

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DGT probes consisted of a 0.08 cm-thick bisacrylamide-crosslinked polyacrylamide diffusive gel,40 overlain with

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a 0.01 cm-thick cellulose nitrate filter membrane (0.45 µm poresize; Millipore). The measurement of Fe(II),

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As(III) and Sb(III) co-distributions in the same sediment location was made possible by using the diffusive gel of

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the mercapto-silica DGT probe for colorimetric Fe(II) DET analysis, as described previously.27, 28


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Deployment and analysis of DGT-DET samplers. DGT-DET sediment probes were deoxygenated for at least

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12 h in 0.01 mol L-1 NaNO3, via sparging with high purity N2 gas to minimize disruption to the anoxic zone of the

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sediment during deployment. One MBL and one mercapto-silica DGT probe were deployed in each mesocosm

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for 24 h on days 3 – 4 (oxic), days 17 – 18 (anoxic) and days 22 – 23 (oxic) (see Figure S1 for probe schematic

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and further deployment details). To minimize the interruption of anoxic conditions during DGT-DET deployment

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and retrieval, the mesocosm lids were removed and the stirrers turned off for no longer than five minutes. While

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probes were gently but rapidly inserted into the sediment, the resuspension of very fine particulates from the

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sediment surface was unavoidable. A stainless steel scalpel was used to slice the gels from the probe exposure

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window immediately upon retrieval; the diffusive gel of the mercapto-silica probes was then analyzed for Fe(II)

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at high-resolution using computer imaging densitometry, as described previously.21,

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concentrations were also determined as described previously (see Table S2 for further details).25

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Solid-phase analysis. Small intact cores of sediment were taken using a 20 mL plastic syringe barrel (18 mm

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internal diameter) from the control and treatment mesocosms on day 18 (anoxic phase) and immediately frozen at

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-20°C for the analysis of solid-phase antimony speciation by X-ray Absorption Spectroscopy (XAS) at the

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Australian Synchrotron. Cores were sliced from 0-1, 1-2, 2-4 and 4-6 cm intervals in an N2-filled glove bag under

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ambient conditions and homogenized. A sub-sample from each depth was mixed with glycerol and frozen for

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transportation to the beamline. Details of the XAS analysis are provided in the Supporting Information (Table

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S2).

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Results and Discussion

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Water column parameters. Dissolved oxygen concentrations remained stable in the control mesocosms at

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~100% saturation (277 ± 2 µmol L-1) throughout the incubation (Figure 1). Fe(II) concentrations remained stable

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as indicated by the low average concentrations (0.43 ± 0.03 µmol L-1) and negligible flux throughout the

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incubation (Table 1). After treatment mesocosms were sealed, dissolved oxygen concentrations steadily

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decreased at a rate of 116 ± 7.74 µmol m-2 h-1 (day 5 – 10, Table 1). Anoxic conditions took 90 h to establish and

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were maintained for ten days (day 10 – 20). Treatment mesocosms were reoxygenated on day 20; a rapid return to

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pre-treatment dissolved oxygen concentrations occurred in less than 24 hours (277.8 µmol L-1). During the anoxic

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phase, there was a significant flux of Fe(II) from the sediment to the water column (14.2 ± 0.91 µmol m-2 h-1),

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Time averaged DGT


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with maximum concentrations reaching 70.1 µmol L-1 in the overlying water. Upon reoxygenation,

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concentrations rapidly decreased within 24 hours to 4.1 µmol L-1 (Table 1). Eh and pH remained stable in the

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control mesocosms throughout the experiment, with an average Eh reading of 545 ± 13 mV and an average pH of

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5.1 ± 0.06. Eh decreased immediately in treatment mesocosms after they were sealed (day 5) and reached a

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minimum of 209 ± 12 mV (day 20). pH increased in treatment mesocosms from 5.1 to 5.9 after anoxic conditions

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were established (0% DO on day 10), which was likely due to increased alkalinity generated by increased Fe(III)-

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and/or sulfate-reduction in the sediment under anoxic conditions.41 The pH peaked at 6.5 just after reoxygenation

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and remained high due to alkalinity remaining in the water column. DOC concentrations remained stable under

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constant oxic conditions in control mesocosms, increasing slightly from 4.10 ± 0.2 mg L-1 on day 3 of the

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incubation to 5.22 ± 0.09 mg L-1 on the final day of the incubation (Figure S2). Similar initial DOC

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concentrations were measured in treatment mesocosms (4.81 ± 0.06 mg L-1); there was a rapid increase under

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anoxic conditions to 16.95 ± 0.88 mg L-1 and a negligible decrease upon reoxygenation (16.56 ± 0.75 mg L-1)

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(Figure S2). The DOC release may be due to the reductive dissolution of iron oxides and/or the increase in pH,

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both of which can affect DOC solubility and desorption.42, 43 Dissolved sulfur concentrations slightly increased in

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control mesocosms (1.79 ± 0.33 mg L-1 on day 3 to 2.20 ± 0.05 mg L-1 on day 23) while treatment mesocosms

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slightly decreased (1.73 ± 0.03 mg L-1 to 1.43 ± 0.45 mg L-1) (Figure S3). This divergent behavior could be

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suggestive of the oxidative dissolution of reduced sulfur during oxic conditions and sulfate reduction during

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anoxic conditions.44 Both treatments showed a sharp increase in concentrations on day 15 in the middle of the

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anoxic period (2.91± 0.99 mg L-1 in control and 2.45 ± 1.19 mg L-1 in treatment) which dropped again two days

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later; most likely due to the disturbance associated with the deployment of DGT-DET probes and removal of

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sediment cores (Figure S3).


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Figure 1. Average (n=3) dissolved water column oxygen, pH, ferrous iron, Eh and metalloid species As(III),

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As(V), Sb(III), and Sb(V) for control ( ) and treatment ( ) mesocosms. Error bars represent ± 1 standard

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deviation of the mean. Dotted lines indicate beginning (day 10) and end (day 20) of water-column anoxia. ACS Paragon Plus Environment

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Under oxic conditions in the control mesocosms, arsenic and antimony species showed opposite trends in water

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column concentrations (Figure 1). Low average concentrations of As(III) (20.8 ± 9.1 nmol) and negligible fluxes

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indicated stability in control mesocosms throughout the incubation (Table 1). Average As(V) concentrations were

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higher (44.4 ± 15.8 nmol), and there was a low but significant increase in water column concentrations over time

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(Table 1), supporting thermodynamic predictions favoring As(V) dominance under oxic conditions.45 This could

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have been the flux of As(V) species to the water column, or As(III) from the sediment which was subsequently

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oxidized to As(V) (Table 1). In contrast, average antimony species concentrations were much higher (241 ± 70.2

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nmol L-1 Sb(V) and 234 ± 108 nmol L-1 Sb(III)) throughout the incubation. There was evidence of both species

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fluxing from the sediment over time, although speciation could have changed in the water column (24.3 ± 6.56

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nmol m-2 h-1 Sb(V) and 32.2 ± 13.7 nmol m-2 h-1 Sb(III)) (Table 1). Furthermore, given that the concentration of

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solid-phase antimony was ~2-fold higher than arsenic, the almost 10-fold higher aqueous concentration of

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antimony compared to arsenic suggested that antimony was more mobile under oxic water-column conditions in

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this system or was present as a more labile species. Higher mobility of antimony compared to arsenic under oxic

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conditions was previously reported under laboratory conditions,46 as well as in natural surface and ground

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waters.9, 47

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Table 1. Sediment- water column fluxes of dissolved oxygen (µ µmol m-2 h-1), Fe(II) (µmol m-2 h-1), As(III)

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(nmol m-2 h-1), As(V) (nmol m-2 h-1), Sb(III) (nmol m-2 h-1) and Sb(V) (nmol m-2 h-1) in the treatment and

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control mesocosms during the anoxic treatment phase. All fluxes were calculated from the slope of the

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linear regression over days 10 – 20, except for DO, which was calculated over days 5 – 10. Negative fluxes

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indicate uptake by the sediment. *significant regression where p

Diffusive Gradients in Thin Films Reveals Differences in Antimony and

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