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Article

The Impacts of Ethanol and Freeze–Thaw Cycling on Arsenic Mobility in a Contaminated Boreal Wetland

by
Joseph Radford
1,2,*,
Kimber E. Munford
1,3,
Nadia Mykytczuk
4 and
Susan Glasauer
1
1
School of Environmental Sciences, University of Guelph, Guelph, ON N1G 2W1, Canada
2
Department of Civil and Mineral Engineering, University of Toronto, Toronto, ON M5S 1A4, Canada
3
Department of Biology, University of Waterloo, Waterloo, ON N2L 3G1, Canada
4
School of Natural Sciences, Laurentian University, Sudbury, ON P3E 2C6, Canada
*
Author to whom correspondence should be addressed.
Soil Syst. 2025, 9(2), 37; https://doi.org/10.3390/soilsystems9020037
Submission received: 7 February 2025 / Revised: 8 April 2025 / Accepted: 15 April 2025 / Published: 21 April 2025
(This article belongs to the Special Issue Soil Bioremediation)
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Abstract

:
Pyrite-bearing waste rock from legacy gold mines is a source of elevated arsenic, sulfate, and iron in the surrounding environments due to leaching. Waste rock in environments that experience cold winters is of particular concern because freeze–thaw cycling may mobilize elements through degradation and release of organic matter and accelerated mineral weathering. In boreal zones, wetlands are common recipients of mine-waste run-off, and microbial processes in wetland soil may promote the retention of mobilized elements, such as arsenic. We investigated the impacts of freeze–thaw cycling and ethanol amendment on soil from an arsenic-contaminated wetland in anoxic microcosms. Ethanol-amended microcosms exhibited enhanced microbial sulfate reduction, leading to sulfide precipitation and increased retention of arsenic in the soil. Sequential extraction studies indicated a shift of arsenic into more stable sulfide-bound fractions. The addition of ethanol significantly increased the growth of Geobacter spp. and other select sulfate-reducing bacteria. Freeze–thaw cycling increased dissolved arsenic over short time periods only and had no detectable impacts on microbial activity. These findings suggest that the use of ethanol as an amendment to wetlands during spring thaw may enhance arsenic sequestration in mining-impacted soils and may provide a viable remediation strategy for cold-climate environments, where seasonal freeze–thaw cycling could otherwise contribute to arsenic mobilization.

1. Introduction

Waste rock and tailings from the processing of metal bearing ores for metals is a key source of contamination at legacy mine sites and mine-waste impoundments. The erosion and weathering of waste minerals mobilize mineral particulates and dissolved constituents, which can be exacerbated by seasonal climate fluctuations [1,2]. In Canada and other cold climate regions around the world, wetland soils are often natural recipients of contaminated tailings runoff [3,4,5]. These wetlands serve as dynamic interfaces between terrestrial and aquatic ecosystems, representing critical systems that influence the transport, transformation, and sequestration of contaminants [5].
An element of major environmental concern in legacy gold mines is arsenic (As), an acutely toxic metal which can concentrate in waste rock (“tailings”) generated during the processing of gold-bearing pyritic ores [6]. Arsenic occurs in arsenopyrite (FeAsS) and other minerals associated with many pyritic ores, where it is liberated during oxidative weathering [7,8]. The oxidation of FeII and SII in pyrite (FeS2) and associated minerals results in the formation of acid mine drainage (AMD), which is characterized by low pH and elevated concentrations of dissolved iron, sulfate, and elements of concern such as arsenic [9,10,11]. Although arsenic can exist in several oxidation states, the oxidation states of As in mine-waste discharge of primary concern are inorganic arsenate (AsV) and arsenite (AsIII). Under acidic and oxidizing conditions, As exists mainly as the AsV oxyanions H2AsO4− and H3AsO4, while acidic and reducing conditions favor the AsIII oxyanion H3AsO3 [12,13].
One important approach to remediating AMD-affected water in soils that receive discharge is based on microbial sulfate reduction. This process is facilitated by sulfate-reducing bacteria (SRB) which couple the oxidation of an electron donor, often organic carbon, to the chemical reduction of sulfate [14]. The reduction of sulfate counters AMD by producing alkalinity, while aqueous sulfide forms complexes with divalent metals that can precipitate as relatively insoluble metal-sulfide minerals [15,16]. In reducing conditions, hydrous AsV can be reduced to hydrous AsIII by aqueous sulfide species (i.e., H2S) or directly by arsenate-respiring microbes [17,18,19]. Iron and sulfide minerals in soil, such as orpiment (As2S3), realgar (As4S4), and pyrite (FeS2), as well as organic material, adsorb hydrous species of AsIII such as H2AsO3− [20,21,22,23]. In mildly acidic (pH 4–6), reducing environments, hydrous AsIII (H2AsO3−) can also precipitate with FeII and/or H2S as Fe-As or S-As minerals, predominantly as arsenopyrite (FeAsS) and orpiment (As2S3) [12,24].
Runoff from mining waste rock is typically deficient in organic substrates. In contrast, a wetland that receives runoff contains abundant complex carbon sources, with simple carbon available from the degradation of plant matter and as by-products of microbial metabolic processes [25,26]. Studies on natural wetlands, constructed wetlands, bioreactors, and lab-scale incubation systems have shown that low molecular weight, labile carbon substrates such as acetate, lactate, and ethanol can accelerate sulfate reduction rates and increase metal removal from solution [27,28,29,30,31,32]. Sulfate-reducing bacteria are capable of utilizing a number of simple or easily biodegradable carbon sources, including acetate, lactate, glucose, ethanol, vegetable oil, and starch [33,34]. The ability to stay active at pH values ranging from 3 to 9.5 and temperatures ranging from −1.8 °C to over 100 °C has allowed for the success of SRB-based remediation systems in many climates [19,35,36,37]. Nevertheless, the effectiveness of SRB-based remediation depends strongly on environmental conditions; although SRBs are adaptable, optimized conditions in cold, northern environments have not been well defined.
The site of the former Long Lake gold mine south of Sudbury, Ontario, Canada is well suited for exploring microbially mediated immobilization of arsenic, iron and sulfate in an AMD-affected wetland that is situated in a cold northern climate zone. An uncontained, legacy tailings area is the source of arsenic-laden leachate to the site, with most runoff discharging to a creek that feeds downstream into a wetland area and the nearby Long Lake Bay. The wetland area may act as a point of sequestration for aqueous sulfate and arsenic through microbially mediated precipitation of metal sulfides, as demonstrated in previous natural and constructed wetland bioremediation studies [25,26,38]. Although there are many studies regarding arsenic mobility in mining environments within temperate climates, there is a gap regarding research in northern temperate and boreal regions. A seasonal pattern of fluctuations between frozen and thawed conditions within the upper layer of soil in early spring, unique to these environments, is of particular interest.
Freeze–thaw cycling during early spring conditions can conceivably mobilize previously sequestered arsenic from mining-impacted wetland soils because of more highly oxidized conditions, reflecting suppressed dissimilatory reduction of sulfate and/or metals. Given that freeze–thaw cycling is expected to increase in some northern environments due to decreased snowpack and increased temperatures during the winter and early spring, it is important to understand the microbial response [39,40]. Adding sources of organic C to the wetland during the sensitive period of early spring could help to sequester elements of concern by stimulating dissimilatory sulfate and arsenic reduction. Ethanol is a low cost and readily available carbon substrate that has been shown to efficiently stimulate biogenic sulfide production [41,42].
The objective of this study is to evaluate the effectiveness of microbially mediated sulfate reduction in controlling arsenic mobility within a cold-climate mining-impacted soil. Specifically, we aim to assess how seasonal freeze–thaw cycles influence arsenic sequestration and release, and whether ethanol amendments can enhance sulfate-reducing bacterial activity to improve arsenic retention. Understanding the dynamics of carbon amendments on microbial activity in the framework of arsenic mobility can help to inform tailings effluent management systems, including actively managed natural or constructed wetlands, in this climate.
This investigation focused on the biogeochemical cycling of arsenic in wetlands at the former Long Lake gold mine near Sudbury, a contaminated legacy site. The former Ministry of Energy, Northern Development and Mines (now Ministry of Mines; MINES) has previously reported the need for remediation of tailings leachate from major tailings area 1 (TA-01) [43].

2. Materials and Methods

2.1. Sample Location and Procedure

The site of the former Long Lake gold mine located south of Sudbury, Ontario, Canada hosts 163,000 m3 of abandoned tailings, with most located in major tailings area TA-01 (46°18′45″ N 81°08′50″ W). This mine was operational from 1908 to 1939 and is now managed by the Ministry of Mines [44]. The tailings generate acid mine drainage containing sulfate, iron, and arsenic that discharge into Long Lake bay after traversing through an intermediary stream named Luke Creek (Figure 1). The top sediment layer of TA-01 tailings were determined to a maximum concentration of 17,567 mg/kg, with an average of 2287 mg/kg in the open tailings and a higher average of 5437 mg/kg along the edge, where tailings have migrated over time [45].
Soil samples were collected from a wetland adjacent to Luke Creek, situated between TA-01 and Long Lake that receives runoff from the tailings through seasonally high water levels and lateral discharge from the creek, which has formed in sandy soil. Soil was taken from within two distinct 2 m by 2 m plots, one located 10 m into the wetland from the bank of the stream, termed the runoff-contaminated wetland (RCW), and another plot located 100 m into the wetland from the bank of the stream, termed back wetland (BW). The BW was used as a comparative site because it does not receive direct inflow of tailings runoff from Luke Creek, although it may receive seepage during high water periods. Four subsamples were retrieved from within each site and combined equally by weight to provide a composite sample for each location. The RCW site contained a surface layer of undecomposed plant detritus ranging from 5 to 15 cm in thickness covered with 20–30 cm of water at the time of sampling (June 2019). Below the undecomposed plant detritus was a soil layer comprising a mixture of decomposed organic material and finer mineral particulates. The BW site had a thicker layer of plant detritus (15–20 cm) with 5–10 cm of water cover. For sampling, the large, loose undecomposed organic matter was removed to reach the decomposed organic matter and mineral soil beneath, identified by a denser and sludge-like consistency. The mineral soil was collected and stored in sealed plastic bags in the dark at 4 °C before use in microcosm studies.
The pH and dissolved O2 (dO2) of the surface soil were determined using the Thermo Scientific Orion Star A325 portable meter (Thermo Fisher Scientific, Waltham, MA, USA). To obtain pH values, soil was analyzed on-site immediately after sampling. The pH probe was inserted into the bags of soil so that the probe was completely covered and allowed to equilibrate with the pore water. This was repeated in triplicate for bags of soil from six individual points within each plot, for a total of 18 measurements per plot. The dO2 measurements were taken from the soil–water interface for all points within the plots, as this was the limit for the depth the probe could reach. The RCW soil had a pH of 5.9–6.0 and the site had a dO2 of 0.1–0.5 mg/L, while the BW soil had a pH of 5.4–5.7 and the site had a dO2 of 7.2–10 mg/L. Since dO2 was measured at a higher depth than the depth at which the soil was extracted, the reported values are likely higher than the actual soil values.

2.2. Freeze–Thaw Experiments

A series of microcosms were set up to evaluate the impacts of freeze–thaw cycling (FT) and ethanol amendment on soil chemistry, microbial respiration, and microbially influenced arsenic chemistry and mobility. Microcosms containing wetland soil were supplemented with Minimal Freshwater Medium [46] amended with comparative treatments of arsenate, ethanol, and sulfate. Treatment labels and constituent concentrations are detailed in Table 1. The microcosms were then subjected to FT cycles to simulate spring-thaw conditions. This approach aimed to assess the potential for arsenic (As) mobilization during spring conditions and to evaluate whether simple chemical amendments could mitigate As loss.
Sodium sulfate and ethanol were added in a 1:1.5 molar ratio (10 mM:15 mM) as previous studies have shown this ratio can achieve optimal sulfate reduction [41]. Ethanol was chosen as the carbon source, over other feasible sources such as lactate or acetate, due to its demonstrated ability to stimulate sulfate-reducing activity and increase sulfate-reducing bacteria (SRB) biomass and its cost-effectiveness for potential large-scale applications [41,42,47,48].
Incubation experiments consisted of two sets of microcosms for each of the above listed treatments: one set subjected to three FT cycles, and another set maintained at 4 °C. Microcosms exposed to freeze–thaw cycling before incubation are denoted by “FT” and treatments not exposed to FT cycling are denoted by “non-FT”. After the FT cycling period, both sets were incubated at room temperature for 19–36 days. A set of 0.5% ethanol-free formaldehyde sterilized incubations were conducted in parallel to the treated microcosms and were used as biological controls. Several different sterilization or bacterial inactivation methods were tested, including multiple rounds of autoclaving and fumigation, but none resulted in complete soil sterilization [49,50,51,52,53]. The formaldehyde treatment was ultimately chosen because heat-, light-, and pressure-related geochemical reactions were avoided and sustained microbial inactivation was demonstrated in previous research [49].
The incubations were conducted in 600 mL borosilicate serum bottles (DWK Life Sciences, Millville, NJ, USA) sealed with rubber stoppers and aluminum crimp tops. Soil was taken from bulk collection bags of the appropriate sample site (RCW or BW) and homogenized in a sterile beaker by mixing with a sterile glass stir bar. Large pieces of organic matter, including twigs or large roots, were discarded, and 50 g of homogenized soil was distributed to each bottle. 250 mL of sterile freshwater minimal medium was added to each bottle and bubbled with sterile N2 gas (0.45 µm antimicrobial gas line filter) for ~5 min before sealing.
For FT cycling, microcosms were placed in a Styrofoam insulated tray that covered the bottles up to the liquid level, simulating natural freezing conditions from the surface downward. They were then placed in an environmental chamber and subjected to three cycles of freezing and thawing between +4 °C and −4 °C, with 24 h intervals per temperature phase. Temperatures in the chamber completed transition in <2 h, but due to the insulating foam, the bottles did not completely freeze or thaw until at least 16 h after the temperature was changed. Following FT cycling, bottles were incubated at room temperature in an anaerobic chamber. Sampling was performed under anaerobic conditions, except for end point soil extraction, which took place in a biosafety cabinet.

2.3. Gas Chromatography

In order to track microbial respiration, the headspace of each microcosm was sampled for carbon dioxide (CO2). Gas extractions were analyzed using the SRI 8610 multiple gas analyzer gas chromatograph (GC) with methanizer, FID, and TCD detection (SRI Instruments, Torrance, CA, USA). The GC separates gasses and detects CO2 using a silica gel packed column and a molecular sieve packed column. Hydrogen was used as a carrier gas at a flow rate of 20 mL/min. The methanizer was set to 380 °C, the valve oven set to 90 °C, the TCD current set to low, and the FID current set to high. Atmospheric air, a 5000 ppm CO2 standard, and a 50,000 ppm CO2 standard were used to create a standard curve for each experiment. Five mL of headspace was extracted from each microcosm using a gas-tight glass syringe and injected directly into the 1 mL sample loop with the excess gas used to flush the line. This procedure was repeated for two time replicates. Results for both time replicates showed similar trends, so only one was reported here. The N2000 chromatography data system (Surwit Technology Inc., Hangzhou, China) was used for data acquisition and analysis.

2.4. Spectrophotometric Sulfate Analysis

Sulfate was analyzed using a modified version of the Hach SulfaVer 4 method, which is equivalent to the USEPA method 375.4 for sulfate testing in wastewater. This is a turbidimetric method in which the SulfaVer 4 reagent (Hach; barium sulfate) reacts with sulfate in the sample to form turbidity proportional to sulfate concentration. This method was developed for a Hach portable colorimeter and was adapted to an Agilent UV-VIS spectrophotometer (Agilent Technologies, Santa Clara, CA, USA) for this study. The same lower detection limit as the Hach colorimeter (Hach Company, Loveland, CO, USA) (~2 mg/L) and a higher upper detection limit (100 mg/L) was achieved. To prepare the samples for analysis, 1 mL of microcosm media was extracted, filtered using a 0.45 µm syringe filter (PES, Fisherbrand Basix, Thermo Fisher Scientific, Waltham, MA, USA), and diluted 10× with deionized water. A SulfaVer 4 barium sulfate packet was added to the diluted sample, shaken until dissolved, and reacted for 5 min. The reacted solution was immediately analyzed in duplicate at 450 nm. If readout values between duplicates exhibited a difference greater than 1%, the analysis for that sample was repeated. Triplicates of a given treatment were averaged and reported with standard deviation. Sulfate standards were prepared with sodium sulfate in deionized water at 2, 5, 10, 20, 50, and 100 mg/L concentrations. The standard error for the spectrophotometric readout was determined to be 5%.

2.5. Sequential and Total Extraction of Wetland Soils

The soil was characterized from the runoff-impacted area of the wetland (RCW) and the back-wetland (BW) site in order to establish the geochemical distribution for the elements of interest. Soil was analyzed for As, Fe, Mn, Ni, and Zn by sequential and total extraction. Sequential extraction is an analytical procedure used to assess the elements associated with operationally defined solid phases. Sequential extraction of arsenic and other cations was conducted using a modified version of the method described by Huang and Kretzschmar [54]. Soil was freeze dried and passed through a 2 mm sieve. 0.25 g of sieved soil was used for each soil or microcosm extraction. Six fractions were targeted: (1) soluble and exchangeable, (2) organic matter, (3) acid-volatile sulfides and very poorly crystalline Fe (hydr)oxides, (4) poorly crystalline Fe (hydr)oxides, (5) sulfides, and (6) residual. The original method was modified to omit the Mn oxides and crystalline Fe (hydr)oxides fractions because the concentrations of manganese in the soil were negligible, and for the purposes of this study, it was not important to separate the crystalline mineral fraction from the residual fraction. A reverse aqua regia digestion was conducted for the combined crystalline Fe (hydr)oxide and residual fraction. For this extraction, 9 mL HNO3 and 3 mL HCl were added to the residual soil in Teflon digestion bombs at room temperature and allowed to dissolve organic matter for 24 h or until the acid no longer reacted with the organic matter. The bombs were sealed and placed in an oven at 120 °C for 16–20 h before being allowed to cool, followed by filtration using Whatman 42 2.5 mm ashless filter papers (Millipore Sigma, Burlington, MA, USA, WHA1442-042). Total extractions were conducted using the same reverse aqua regia digestion procedure.
The first extraction of the RCW and BW site soil was conducted after sampling, without incubation, in 12 replicates each. Montana soil II is a standard reference material certified by the National Institute of Standards and Technology and was included alongside extractions. Subsequent extractions of post-incubation amended RCW soil were conducted once on each microcosm triplicate, with results averaged for each treatment and reported with standard deviation.

2.6. Arsenic Speciation

The oxidation state of aqueous arsenic was determined using HPLC-ICP-MS (Agilent 1260 Infinity and Agilent 7700x, Agilent Technologies, Santa Clara, CA, USA) for inorganic AsIII and AsV at the Biotron facility at Western University in London, ON, Canada. The method reported an R2 of >0.998. An amount of 5 mL of liquid was extracted through the sealed rubber top using a disposable sterile syringe, filtered through a 0.45 µm syringe filter, and diluted to 15 mL using deionized water. The redox state of the inorganic arsenic species was preserved with 500 ppm EDTA and samples were stored at 4 °C until analysis.

2.7. Determination of Metal Concentrations

The Varian Vista Pro CCD Simultaneous ICP-OES with an axial viewed plasma was used for the determination of arsenic and iron in both liquid extracts and sequential/total extraction solutions. Potassium and calcium were also analyzed using this method; however, they were found to be of limited significance to this study and are therefore not reported. Analysis was conducted by Peter Smith at the University of Guelph, Guelph ON, Canada. Standards were created by diluting ICP-MS 1000 mg/L stock solutions (SCP Scientific, Mississauga, ON, Canada) to working analytical range. All liquid extracts from microcosms were syringe filtered (0.45 µm, PES, Fisherbrand Basix), acidified using 2% HNO3, and stored at 4 °C until analysis. All elements were analyzed in triplicate. Results were reported as either a range of values or as an average of the values with the standard deviation.

2.8. DNA Extraction and 16S rRNA Gene Analyses

0.25 g of the pre-incubated and post-incubated RCW soils were extracted in triplicate using the Qiagen DNeasy Power Soil Pro kit, following the manufacturer’s instructions (Qiagen GmbH, Hilden, Germany). The quantity and quality of the extracted DNA was inspected using the Nanodrop spectrophotomer ND 2000 (Nano-Drop Technologies, Wilmington, DE, USA). Genomic DNA was amplified by PCR using primers targeting the 16S rRNA gene of bacteria and archaea. PCR was conducted following these steps: (i) an initial denaturing step at 95 °C for 5 min, (ii) 30 cycles of 95 °C for 30 s, 58 °C for 30 s, 72 °C for 60 s, and (iii) a final extension step of 72 °C for 5 min. PCR amplicons were then sequenced using paired-end MiSeq Illumina Sequencing at Metagenom Bio Inc. (Waterloo, ON, Canada).
Demultiplexed sequences were processed using DADA2 v1.8 [55] managed through QIIME 2 v.2019.7 [56]. In this workflow forward and reverse reads were truncated at decreasing quality, primers were removed, and paired reads were assembled after Illumina sequencing error modeling and correction. Subsequently, chimeric ASVs were removed by reconstruction against more abundant parent ASVs. An amplified sequence variant (ASV) table was then constructed for downstream analysis. An ASV table is analogous to an OTU table but constructed through Illumina denoising rather than clustering techniques. A rarefaction curve was generated prior to downstream filtering of data, indicating adequate sequencing depth for all samples (Supplementary Figure S1). Taxonomy was assigned to representative sequences using a naive Bayesian classifier implemented in QIIME 2 with scikit-learn (v.0.21.3) trained against SILVA release 134 clustered at 99% identity. Assignments were accepted above a 0.7 confidence threshold.
Data analysis for 16S rRNA data was performed using the R programming language, version 3.17 [57], utilizing the packages “vegan”, “microeco”, and “phyloseq” for all analyses [58,59,60]. 413,014 total 16S rRNA sequences were generated by sequencing. Organelles were trimmed from data and any ASVs unassigned at the Kingdom level were removed prior to downstream analysis, resulting in 412,131 reads. To evaluate differences in microbial communities among samples, bar graphs of relative abundance were generated using “microeco”. Differences among treatments (amendments, freeze–thaw) were evaluated at the family level using permutational analysis of variance (PERMANOVA) in “vegan”. Finally, ANOVA-like differential expression was used to evaluate differences in abundance of microbial taxa between ethanol-amended microcosms and those without ethanol [61].

3. Results and Discussion

3.1. Distribution of Arsenic and Iron in Wetland Soil Impacted by Mine Water Discharge

Two subsites of the wetland adjacent to Luke Creek were assessed; one was immediately adjacent to the creek (runoff-contaminated wetland, or RCW), and the other was approximately 100 m from the creek, within the wetland (back wetland, or BW). Prior data from the former Ministry of Northern Development and Mines (2017) indicated that arsenic- and iron-laden runoff from the upstream tailings area TA-01 feeds into Luke Creek, which deposits metals and arsenic into the downstream wetland [62]. Results from the acid digests corroborated those findings, revealing elevated concentrations of As and Fe in the wetland soil traversed by Luke Creek. The RCW soil had significantly higher average concentrations of arsenic (13,479 ± 1055 mg/kg) and iron (47,268 ± 3695 mg/kg) than BW soil (134 ± 75 mg/kg As, 6842 ± 1152 mg/kg Fe). By comparison, average As concentrations in RCW soil are significantly higher than TA-01 tailings, which was previously reported to have average concentrations between 2287 mg/kg (open tailings) and 5437 mg/kg (tailings edge) [45]. The concentrations of arsenic in the BW soil suggest that arsenic-laden runoff or seepage migrates beyond the areas immediately proximate to TA-01 or Luke Creek (see Figure 1); alternatively, these elevated concentrations could reflect naturally high background levels of arsenic in the BW soil. While mean arsenic concentrations in several uncontaminated Canadian soil types have been reported in the range of 4.8–13.6 mg/kg, natural background concentrations can range by orders of magnitude [63]. Nevertheless, arsenic concentrations found at both wetland sites are markedly higher than Canadian soil quality guidelines (12 mg/kg), which poses a serious ecological threat if present in bioavailable soil fractions [63].
In order to assess the chemical and mineralogical environment of As in the soil, a sequential extraction method was used to classify arsenic and iron into seven operationally distinct pools. Studies have shown that arsenic adsorbed to, or bound within, organic matter, poorly crystalline iron hydroxides, or acid-volatile sulfides is subject to remobilization by oxidation from iron- or sulfate-oxidizing bacteria, exposure to oxygen, or decreases in pH [20,64,65]. The distribution of arsenic and iron can be expressed in terms of increasing extractant reactivity for the seven fractions (Table 2). The concentrations of As and Fe in the most mobile fraction (soluble and exchangeable elements) are low, relative to most of the other fractions, suggesting that sorbed As is not the major pool, although this fraction may vary seasonally.
The organic fraction contained approximately 17% of total arsenic and iron. The bulk of As and Fe were contained in the acid-volatile sulfides fraction and in the poorly crystalline Fe/Al fraction, together representing about 60% of arsenic and iron in the RCW soil. This supports that poorly crystalline iron and sulfide minerals are the major solid phases for arsenic.
Acid-volatile sulfides (AVS) are a group of minerals that encompasses poorly crystalline, fine grained metal sulfide phases [66]. The largest proportion of iron was associated with the AVS (Fraction 3), at 43.55%. The sulfide mineral fraction (Fraction 5) represents arsenic in less reactive sulfide minerals including arsenopyrite (FeAsS) and amorphous orpiment (As2S3); this fraction made up 18.3% of total As. Arsenic and iron are strongly correlated in Fractions 3, 4, and 5 (r = 0.948, p < 0.01). Conversely, a relatively low amount of arsenic and iron exists in the most refractory mineral forms, including highly crystalline and silicate-associated forms (1.46% total As, 6.04% total Fe in Fractions 6 and 7). Arsenic and iron also showed a weak and non-significant correlation in these fractions (r = 0.527, p > 0.1). Together, the data suggests that most arsenic is associated with the soil in relatively reactive mineral phases and therefore has strong potential for remobilization in conditions of redox fluctuation.

3.2. Changes in the Distribution of As and Fe After FT Cycling

Sequential extraction was used to investigate changes in the distribution of arsenic and iron among different solid fractions between pre-incubation bulk soil (Table 2) and post-incubation microcosm soils. Due to the natural heterogeneity of RCW soil, total concentrations of arsenic and iron varied significantly between microcosms (up to 15%). The total arsenic and iron concentrations also differ between pre-incubation bulk soil (12,994 mg/kg As; 47,268 mg/kg Fe) and post-incubation microcosm soil (7139–9939 mg/kg As; 29,168–39,011 mg/kg Fe), supporting that As and Fe associated with the solids were mobilized during the incubation. To control for this, arsenic and iron in each fraction were expressed as a percentage of the total within their respective microcosm, with post-incubation changes in soil fractions reported as ΔAs and ΔFe (Figure 2).
For all treatments, there was very little change in Fe and As for Fraction 1, the soluble and exchangeable fraction, indicating that these elements were mobilized from the less reactive fractions. After incubation, there was a clear decrease in arsenic in fractions associated with organic matter (Fraction 2) and poorly crystalline Fe-(hydr)oxides (Fraction 4), and an increase in fractions associated with sulfides (Fractions 3 and 5). The trends for Fe were similar to those for As: decreases in Fractions 2 and 4 and increases in Fraction 3, but the changes in concentrations were generally lesser in magnitude.
For the ETH treatment, the sulfide fraction (Fraction 5) increased significantly in As but decreased in Fe—this suggests that arsenic was sequestered by the existing sulfides in the RCW soil. The ARS and SUF treatments showed a tendency to increase in As and Fe for Fraction 5, although the variability was high and the magnitude of the difference for Fe was low (Figure 2).
Overall, the microcosms amended with ethanol (ETH) showed the most consistent and substantial changes in arsenic concentration. A greater loss of arsenic and iron in the organic fraction and the poorly crystalline oxide fraction by the end of the incubation in ETH microcosms may be explained by the sulfide-driven chemical reduction of surface bound and coprecipitated FeIII and AsV. Previous research shows that dissolved sulfide can drive both FeIII and AsV reduction in soils. It is, therefore, reasonable to suggest that the production of sulfide in ETH-microcosms promoted the mobilization of these elements from the organic fraction [17,47,67]. Once dissolved, aqueous arsenic and iron can co-precipitate as acid-volatile sulfides or oxides, explaining the greater increase in iron and arsenic in Fraction 5 of ETH microcosms. All treatments showed increases in As and Fe in Fraction 3, but the increase was significantly larger for the ETH microcosms. For most fractions, the differences in ΔAs and ΔFe between FT and non-FT microcosms of the same treatment were either not significant or were relatively small (i.e., less than 1%), indicating freeze–thaw cycling under the conditions of the incubation did not affect As and Fe.
Although some triplicate samples varied significantly in ΔAs and ΔFe values, the two elements were positively correlated under most conditions. To test this, Spearman’s correlation and Spearman’s rank significance tables were used. Correlation patterns were examined across the full dataset (Figure 3a), separated by treatment (Figure 3b), and separated by individual soil fractions (Figure 3c). Overall, a strong significant positive correlation (rs = 0.891, p < 0.01) was found between ΔAs and ΔFe. Fractions 2, 3, and 4 (rs = 0.932, 0.984, and 0.942, respectively), representing the organic, poorly crystalline iron, and acid-volatile sulfide associated fractions, showed significant, very strong positive correlations between ΔAs and ΔFe. Fractions 1, 5, and 6, (rs = 0.750, 0.645, and 0.628, respectively) representing the soluble, sulfide, and crystalline iron associated fractions, showed significant but weak positive correlations for these two elements. The ΔAs and ΔFe in ethanol-free treatments ARS and SUF (rs = 0.891 and 0.930) were correlated more strongly than ethanol-containing treatment ETH (rs = 0.809); ethanol amended microcosms tended to cluster closer than ethanol-free microcosms.

3.3. The Effects of Freeze–Thaw Cycling and Carbon Amendment on Microbial Activity and Water Chemistry

The impact of freeze–thaw cycling on aqueous arsenic, iron, and sulfate concentrations was minimal compared to the influence of treatment type. Microbial respiration was stimulated by ethanol addition, as shown by elevated CO2 concentrations by the end of the incubation period (Supplementary Figure S2). In contrast, freeze–thaw exposure had no significant effect on respiration. Additionally, most of the FT-imposed changes in water chemistry occurred during the establishment of the microcosms and were negligible by the end of the incubation period.
In all non-sterile treatments receiving amendments, increases in dissolved arsenic were observed emerging within the first few days in the ETH treatments, and during the second week of incubation in the ARS and SUF treatments. Only the ETH treatments subsequently declined in concentration before increasing again between days 19 and 36 (Figure 4). After 36 days, all of the treatments had similar concentrations of dissolved As. Dissolved Fe increased in all treatments, with the highest rate observed in ETH; after 19 days, Fe decreased in the ETH microcosms by the end of the incubation period. Dissolved As, and especially Fe, were the highest for the treatments that did not experience freezing (N-ETH). Sulfate declined only in the ETH treatments, beginning between days 4 and 12, accompanied by the formation of black precipitates commonly related to the formation of metal sulfides [68,69]. Concentrations of dissolved sulfate in treatments that did not receive ethanol showed no decrease over the incubation period, and no precipitates were observed. There was no change in sulfate concentrations for the “killed” controls with formaldehyde, supporting that sulfate reduction was inactive.
The increases in dissolved As and Fe for all treatments indicates a dissolution from loosely bound soil fractions, a result that is supported by decreases in concentrations for both metals in soil fractions 1 and 2 (Figure 2). The largest initial increase in dissolved As and Fe seen in ETH treated microcosms is also consistent with the largest decreases in ETH soil fractions 1 and 2. While dissolved As concentrations in the ETH microcosms plateaued and subsequently decreased over the course of incubation—eventually reaching concentrations similar to those in the ARS and SUF treatments—the ETH microcosms still exhibited the greatest net loss of As from the loosely bound sediment fractions. This trend suggests that the early increase in As was primarily due to microbial reduction of soil Fe-oxides, followed by the incorporation of As into AsS precipitates.
The increase in dissolved Fe under the experimental conditions further supports the occurrence of dissimilatory reduction of Fe(III)-oxides to Fe2+ [70]. Arsenic has been shown to readily adsorb to Fe(III)-oxides, and reduction of Fe3+ in these minerals can release sorbed As into solution [71]. The broader trends for sulfate and Fe in ETH microcosms align with the observed timeline of sulfate reduction and the formation of black sulfide precipitates. Biologically induced sulfide production can scavenge As from solution through incorporation into FeAsS minerals, or by surface adsorption, as seen in similar microcosm studies [16,17,72,73]. Similar trends were reported in Saunders et al. (2008), where the injection of a carbon substrate into a contaminated aquifer resulted in an initial increase in aqueous iron and arsenic, followed by a decrease linked to biogenic sulfide precipitation [74].
Non-FT ETH microcosms exhibited slightly more sulfate removal than FT-ETH microcosms by the completion of the incubation, although this trend was weaker in a replicate experiment. We also observed reversed trends for dissolved As and Fe with respect to freeze–thaw exposure in the replicate. These data indicate that the native microbial community in the soil was resilient against freeze–thaw cycling, and that dissolved As, Fe, and sulfate were not significantly impacted.
Samples from three time points (days 1, 12, and 30) from treatments SUF and ETH were also assessed for AsIII and AsV species to investigate the impact of ethanol on the As oxidation state (Figure 5). Concentrations of AsV for all treated microcosms taken 24 h post incubation (Day 1) were between 39.2 and 58.1 mg/L, a decrease from the concentration after the addition of the AsV amendment (~75 mg/L; 1 mM). This initial loss from solution is presumed to be mainly from adsorption or complexation of AsV with organic matter and Fe-oxyhydroxides, both of which are abundant in RCW soil. Organic As was not assessed separately for this study; soluble organic-arsenic complexes were, therefore, not included in the dissolved AsIII and AsV values. Organic chelation of As also supports the observation that total dissolved As increased during incubation across SUF and ETH microcosms (Figure 4) despite the decrease in the sum of inorganic AsIII and AsV (Figure 5).
Despite identical concentrations of the initial amendments, AsV values on Day 1 were higher for FT-SUF and FT-ETH microcosms (54.8 ± 3.3 mg/L), compared to their non-FT counterparts (41.8 ± 6.7 mg/L). A previous study showed that FT cycling of acidic organic rich soils results in the degradation of high molecular weight organic ligands and the release of surface bound trace element complexes, including arsenic [75]. This suggests that FT cycling may have promoted solubilization of AsV from soil prior to amendment, while the degradation of organic matter may have decreased the capacity of the soil to adsorb amended AsV in the first 24 h.
Overall, a negative correlation was observed between the concentrations of AsIII and AsV in all treatments. The largest decrease in AsV was observed between days 1–12 (41–60%) and indicates the establishment of reducing conditions. Increases in AsIII did not account for the loss of AsV, suggesting that some AsIII was lost via complexation with organic matter or adsorption to existing Fe or S minerals. Additionally, greater decreases of AsV seen in ETH microcosms may be a result of microbial reduction to AsIII and rapid incorporation into newly formed sulfide mineral precipitates, as seen in similar incubation studies [17,47] and also aligns with the beginning of sulfate loss (Figure 4). Smaller decreases in AsV were observed between days 12–30, and less than 3 mg/L could not be explained by an increase in AsIII.
The pH for all treatments increased from between 5.8 and 6.1 to 6.8–7.2 over the first 30 days of incubation (Supplementary Table S1). Microcosms amended with AsV, SO42−, and ethanol exhibited the largest increases in pH (ETH, +1.05–1.17) compared to those with only AsV (ARS, +0.68–0.98) or AsV + SO42− (SUF, +0.89–0.99). Bacterially mediated reduction of sulfate consumes protons and produce alkalinity; the precipitation of FeII with sulfide can also increase the pH [68,76]. The greatest pH increase between sampling points occurred in the first six days in the ETH microcosms, which indicates that biostimulation can help to counter AMD and supports that the microbial response was vigorous. The release of acid- and metal-enriched water in spring due to suppressed SRB activity could be ameliorated by adding a carbon amendment to stimulate soil microbes [28,48].

3.4. Microbial Communities

Microbial communities were assessed at the end of the experiment. Experimental conditions alone significantly impacted microbial communities at the family level relative to Time 0 samples (PERMANOVA: F(1,28) = 4.34, r2 = 0.134, p < 0.001). Freeze–thaw cycling did not result in a significant difference in microbial communities (PERMANOVA: F(1,10) = 0.639, r2 = 0.060, p = 0.848), but amendment addition did (PERMANOVA: F(2,9) = 3.81, r2 = 0.459, p < 0.001). Results show notable similarities in abundance data at the order level between ARS and SUF treatments (Figure 6a). Conversely, ethanol addition appeared to have a marked effect on microbial community structure compared to field samples and across both temperature exposures (PERMANOVA: F(1,10) = 5.42, r2 = 0.352, p = 0.0014). Differential abundance testing further revealed that members of the order Desulfuromonadales significantly dominated ETH-amended microcosms (Supplementary Table S2). The most relatively abundant genus belonging to Desulfuromonadales, Geobacter, had the greatest increase in relative abundance in ethanol-amended microcosms, ranging from 19 to 26% in ETH microcosms compared to 3–5% in ARS and SUF microcosms (Figure 6b). Geobacter is a well-studied metal-reducing genus with species capable of using sulfate, iron, and a number of other metals as terminal electron acceptors for anaerobic respiration [77,78,79].
Geobacter spp. were also detected consistently in fresh soil from the field site, with relative abundance in the range of 2–35%. Environmental studies tend to link Geobacter spp. to dissimilatory metal reduction rather than to dissimilatory sulfate reduction [80,81,82,83] suggesting that the link between Geobacter spp. enrichment and sulfate removal in ETH microcosms is through a secondary reaction. Geobacter spp. may have reduced FeIII (hydr)oxides originating in the wetland soil, followed by a secondary reaction between FeII and SO42−. As the sequential extraction method used in this study is unspecific to the redox state, it is possible that SO42− was removed from the solution as Fe-sulfate or Fe-sulfide mineral precipitates, although the presence of black precipitates suggests the latter. Additionally, some Geobacter spp. have been shown to reduce elemental sulfur; a pathway involving initial reduction of SO42− to S0 by a separate SRB and then subsequently to H2S by Geobacter spp. is, therefore, also possible [25].
Bacterial amplicon sequence variants (ASVs) from other sulfate-reducing genera commonly isolated from acid mine drainage, freshwater soil, and constructed wetlands were enriched in ETH microcosms. These included Desulfosporosinus spp. with relative abundances of 3.2–6.8% in ETH compared to 1.3–1.7% in ARS and SUF; Desulfovibrio spp. with relative abundances of 2.2–4.8% in ETH compared to <0.7% in ARS and SUF; and Desulfatirhabdium spp. with relative abundances of 1.0–1.9% in ETH and 0.18–2.67% in ARS and SUF [84,85,86,87].
Understanding shifts in the microbial community within lab-bench microcosms is an important initial qualitative assessment tool when considering the application of amendments to a large-scale system such as a treatment wetland. Together with geochemical data presented in this study, the 16S rRNA gene community profiles indicate that the addition of ethanol stimulated the growth of bacteria commonly associated with sulfate and iron reduction, ultimately resulting in the removal of sulfate and arsenic from solution.
Despite the significant enrichment of certain bacterial groups, the sustained presence of bacterial diversity indicates that the addition of ethanol in ETH microcosms did not trigger the formation of sulfate- or iron-reducing monocultures, even though these species were the drivers of dominant biogeochemical cycling in the soil.

4. Conclusions

Remobilization of previously sequestered metals is a threat to the long-term stability of tailings contamination in many legacy mining environments, including the former Long Lake gold mine. Our study found extremely high concentrations of arsenic and iron in wetland soil influenced by tailings-contaminated runoff from an adjacent creek. Importantly, soil extractions highlighted the risk for remobilization of a substantial proportion of the immobilized arsenic via proton-promoted dissolution from organic matter, poorly crystalline iron-hydroxides, and acid-volatile sulfides. On the basis of the six operationally defined fractions examined, iron and arsenic were significantly correlated. To reduce the risk of oxidative As remobilization from amorphous and loosely bound soil fractions it is suggested that, based on previous research, reducing conditions be maintained at the soil–water interface [88,89].
Lab scale microcosm incubations revealed that amendment with ethanol impacted soil microbiology and geochemistry the most, over time scales of weeks, whereas FT cycling had little impact. At a 1.5:1 ethanol/SO42− amended ratio the microbial community was able to recover from three freeze–thaw cycles and remove ~80% SO42− from solution in just over 1 month. Microcosms with sulfate reduction exhibited a decrease in aqueous iron and arsenic that is posited to be a result of co-precipitation with or adsorption to microbially generated sulfide. Ethanol amendment also resulted in a shift in the 16S rRNA identifiable bacterial community, most notably drastic increases in the abundance of iron-reducing Geobacter spp. and sulfate-reducing Desulfosporosinus spp. These results, combined with the relatively low-cost of ethanol compared to other commonly studied labile carbon additives (i.e., acetate, lactate), suggest that using ethanol to stimulate sulfate- and iron-reducing activity has the potential to simulate arsenic sequestration in contaminated, cold-climate soils.
Conversely, three cycles of FT prior to incubation were found to have little to no impact on the composition of the RCW microbial community and its capacity for respiration and sulfate reduction after several weeks when soil geochemistry indicated recovery from FT exposure. The observations made in this study show that FT did not significantly influence soil arsenic and iron, but this does not exclude the possibility that arsenic may be released immediately after FT, a particularly sensitive time when microbial activity is relatively low, as supported by microbial respiration.
In summary, our findings suggest that the RCW site may be a viable target for further in situ bioremediation of contaminated tailings effluent. The wetland soil hosts a bacterial community susceptible to short-term inhibition by FT cycling but capable of surviving extremely high concentrations of arsenic and functionally resistant to the long-term effects of FT cycling. More research should be performed to investigate shifts in microbial community composition during key seasonal changes (i.e., during spring thaw) for a better understanding of microbial activity after prolonged cold periods. To evaluate the feasibility of field-scale ethanol amendments for enhancing in situ arsenic sequestration, future studies should focus on quantifying arsenic removal rates and identifying optimal ethanol concentrations needed to maximize SRB-driven arsenic immobilization during sensitive spring flush conditions.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/soilsystems9020037/s1, Table S1: Average pH of RCW sediment microcosms during first 30 days of incubation; Table S2: Key taxa which were significantly different (using ALDEx2) in abundance in ETH-amended microcosms at the end of a 30-day incubation; Figure S1: Rarefaction curve for microbial samples generated prior to trimming; Figure S2: Anaerobic CO2 production during a 36-day incubation of amended runoff-contaminated wetland soil (RCW) microcosms. Abbreviations: FT—freeze-thaw exposed; nFT—no exposure to freeze-thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM SO4; ETH—1mM arsenate + 10 mM SO4 + 15 mM ethanol.

Author Contributions

Conceptualization, J.R. and S.G.; Data curation, J.R. and K.E.M.; Formal analysis, J.R., K.E.M. and S.G.; Funding acquisition, S.G.; Investigation, J.R. and N.M.; Methodology, J.R. and S.G.; Project administration, S.G.; Resources, S.G.; Software, K.E.M.; Supervision, N.M. and S.G.; Validation, J.R.; Visualization, J.R. and K.E.M.; Writing—original draft, J.R. and S.G.; Writing—review and editing, J.R., K.E.M., N.M. and S.G. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the Natural Sciences and Engineering Research Council of Canada (NSERC) Discovery Grant to Susan Glasauer.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data supporting the findings of this study are available upon reasonable request from the corresponding author, J.R.

Acknowledgments

We would like to thank Kazuhito Mizutani for extensive help with field sampling and sample management, as well as Peter Smith and James Longstaffe for aiding in technical analysis.

Conflicts of Interest

The authors declare no conflicts of interest.

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Figure 1. Satellite imagery of the Major Tailings Area 1 (TA-01) at the former Long Lake gold mine. Luke Creek transports arsenic, iron, and sulfate-enriched tailings runoff into Long Lake bay and, to a lesser extent, the proximate wetland. Abbreviations: RCW—Runoff-contaminated wetland. BW—Back wetland.
Figure 1. Satellite imagery of the Major Tailings Area 1 (TA-01) at the former Long Lake gold mine. Luke Creek transports arsenic, iron, and sulfate-enriched tailings runoff into Long Lake bay and, to a lesser extent, the proximate wetland. Abbreviations: RCW—Runoff-contaminated wetland. BW—Back wetland.
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Figure 2. Percent change in arsenic (As) and iron (Fe) fractions in run-off-contaminated wetland (RCW) soil after microcosm incubation. Values represent the percentage change in each fraction’s contribution to total As and Fe between pre-incubation and post-incubation conditions. Positive values indicate an increase in the fraction’s contribution to total Fe or As after incubation, while negative values indicate a decrease. Error bars represent standard deviation. Treatments: ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM SO4; ETH—1 mM arsenate + 10 mM SO4 + 15 mM ethanol. Fraction descriptions: (1) soluble and exchangeable, (2) organic-associated, (3) acid-volatile sulfides and very poorly crystalline iron (hydr)oxides, (4) poorly crystalline iron (hydr)oxides, (5) sulfides, (6) crystalline iron minerals.
Figure 2. Percent change in arsenic (As) and iron (Fe) fractions in run-off-contaminated wetland (RCW) soil after microcosm incubation. Values represent the percentage change in each fraction’s contribution to total As and Fe between pre-incubation and post-incubation conditions. Positive values indicate an increase in the fraction’s contribution to total Fe or As after incubation, while negative values indicate a decrease. Error bars represent standard deviation. Treatments: ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM SO4; ETH—1 mM arsenate + 10 mM SO4 + 15 mM ethanol. Fraction descriptions: (1) soluble and exchangeable, (2) organic-associated, (3) acid-volatile sulfides and very poorly crystalline iron (hydr)oxides, (4) poorly crystalline iron (hydr)oxides, (5) sulfides, (6) crystalline iron minerals.
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Figure 3. Correlation plots comparing the percentage change in fractional iron (ΔFe) and arsenic (ΔAs) concentrations between pre- and post-incubation soil. Each point represents changes in an individual sample fraction: (a) all data points across treatments and fractions, (b) data points by treatment type, (c) data points by soil fraction. Treatments: ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM sulfate; ETH—1 mM arsenate + 10 mM sulfate + 15 mM ethanol. Fraction descriptions: (1) soluble and exchangeable, (2) organic-associated, (3) acid-volatile sulfides and very poorly crystalline iron (hydr)oxides, (4) poorly crystalline iron (hydr)oxides, (5) sulfides, (6) crystalline iron minerals.
Figure 3. Correlation plots comparing the percentage change in fractional iron (ΔFe) and arsenic (ΔAs) concentrations between pre- and post-incubation soil. Each point represents changes in an individual sample fraction: (a) all data points across treatments and fractions, (b) data points by treatment type, (c) data points by soil fraction. Treatments: ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM sulfate; ETH—1 mM arsenate + 10 mM sulfate + 15 mM ethanol. Fraction descriptions: (1) soluble and exchangeable, (2) organic-associated, (3) acid-volatile sulfides and very poorly crystalline iron (hydr)oxides, (4) poorly crystalline iron (hydr)oxides, (5) sulfides, (6) crystalline iron minerals.
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Figure 4. Aqueous concentrations of sulfate, arsenic, and iron in runoff-contaminated wetland (RCW) soil microcosms under different treatment conditions. Error bars represent standard deviation. Treatments: FT—freeze–thaw exposed; nFT—no exposure to freeze–thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM sulfate; ETH—1 mM arsenate + 10 mM sulfate + 15 mM ethanol.
Figure 4. Aqueous concentrations of sulfate, arsenic, and iron in runoff-contaminated wetland (RCW) soil microcosms under different treatment conditions. Error bars represent standard deviation. Treatments: FT—freeze–thaw exposed; nFT—no exposure to freeze–thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM sulfate; ETH—1 mM arsenate + 10 mM sulfate + 15 mM ethanol.
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Figure 5. Aqueous inorganic As(III) and As(V) concentrations in amended microcosms at three points during a 30-day incubation. Error bars represent standard deviation. Treatments: FT—freeze–thaw exposed; nFT—no exposure to freeze–thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM sulfate; ETH—1 mM arsenate + 10 mM sulfate + 15 mM ethanol; con—unamended control.
Figure 5. Aqueous inorganic As(III) and As(V) concentrations in amended microcosms at three points during a 30-day incubation. Error bars represent standard deviation. Treatments: FT—freeze–thaw exposed; nFT—no exposure to freeze–thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM sulfate; ETH—1 mM arsenate + 10 mM sulfate + 15 mM ethanol; con—unamended control.
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Figure 6. Relative abundance barplots at the (a) Order and (b) Genus levels of communities from pre- and post-incubated runoff-contaminated wetland (RCW) site sediments, determined from 16S rRNA sequences. Time 0 represents pre-incubated sediment, while all others were extracted post-incubation. Treatments: FT—freeze–thaw exposed; nFT—no exposure to freeze–thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM SO4; ETH—1 mM arsenate + 10 mM SO4 + 15 mM ethanol. Phyla representing <0.5% relative abundance in a single sample and orders representing <2% relative abundance in a single sample were trimmed prior to plotting.
Figure 6. Relative abundance barplots at the (a) Order and (b) Genus levels of communities from pre- and post-incubated runoff-contaminated wetland (RCW) site sediments, determined from 16S rRNA sequences. Time 0 represents pre-incubated sediment, while all others were extracted post-incubation. Treatments: FT—freeze–thaw exposed; nFT—no exposure to freeze–thaw cycling; ARS—1 mM arsenate; SUF—1 mM arsenate + 10 mM SO4; ETH—1 mM arsenate + 10 mM SO4 + 15 mM ethanol. Phyla representing <0.5% relative abundance in a single sample and orders representing <2% relative abundance in a single sample were trimmed prior to plotting.
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Table 1. Arsenate, sulfate, and ethanol amendment concentrations for microcosm treatments.
Table 1. Arsenate, sulfate, and ethanol amendment concentrations for microcosm treatments.
ConstituentsSample Codes and Constituent Concentrations 1
ARSSUFETH
Sodium arsenate1 mM1 mM1 mM
Sodium sulfate-10 mM10 mM
Ethanol--15 mM
1 ARS, SUF, and ETH are short codes representing the specific combinations of sodium arsenate, sodium sulfate, and ethanol used in the three microcosm experiment treatments.
Table 2. Distribution of As and Fe in operationally defined fractions of runoff-contaminated wetland (RCW) soil.
Table 2. Distribution of As and Fe in operationally defined fractions of runoff-contaminated wetland (RCW) soil.
FractionAs (mg/kg)Avg. As (%)Fe (mg/kg)Avg. Fe (%)
Soluble and exchangeable430 (±14)3.30%70 (±2.7)0.15%
Organic2195 (±119)16.89%8425 (±452)17.82%
AV sulfides *, VPC Fe/Al **4674 (±486)35.97%20,584 (±1721)43.55%
Poorly crystalline Fe/Al3127(±165)24.07%8514 (±279)18.01%
Sulfides2380 (±225)18.31%6820 (±1029)14.43%
Crystalline Fe/Al189 (±46)1.46%1461 (±143)3.09%
Residual0 (±0)0.00%1394 (±68)2.95%
Total (sum of fractions)12,994100.00%47,268100.00%
* AV—acid-volatile. ** VPC—very poorly crystalline.
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MDPI and ACS Style

Radford, J.; Munford, K.E.; Mykytczuk, N.; Glasauer, S. The Impacts of Ethanol and Freeze–Thaw Cycling on Arsenic Mobility in a Contaminated Boreal Wetland. Soil Syst. 2025, 9, 37. https://doi.org/10.3390/soilsystems9020037

AMA Style

Radford J, Munford KE, Mykytczuk N, Glasauer S. The Impacts of Ethanol and Freeze–Thaw Cycling on Arsenic Mobility in a Contaminated Boreal Wetland. Soil Systems. 2025; 9(2):37. https://doi.org/10.3390/soilsystems9020037

Chicago/Turabian Style

Radford, Joseph, Kimber E. Munford, Nadia Mykytczuk, and Susan Glasauer. 2025. "The Impacts of Ethanol and Freeze–Thaw Cycling on Arsenic Mobility in a Contaminated Boreal Wetland" Soil Systems 9, no. 2: 37. https://doi.org/10.3390/soilsystems9020037

APA Style

Radford, J., Munford, K. E., Mykytczuk, N., & Glasauer, S. (2025). The Impacts of Ethanol and Freeze–Thaw Cycling on Arsenic Mobility in a Contaminated Boreal Wetland. Soil Systems, 9(2), 37. https://doi.org/10.3390/soilsystems9020037

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