Science of the Total Environment 526 (2015) 215–221

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Impact of mitigation strategies on acid sulfate soil chemistry and microbial community Xiaofen Wu a, Pekka Sten b, Sten Engblom c, Pawel Nowak d, Peter Österholm e, Mark Dopson a,⁎ a

Centre for Ecology and Evolution in Microbial Model Systems (EEMiS), Linnaeus University, Kalmar, Sweden Energy & Environmental Technology, Vaasa University of Applied Sciences, Vaasa, Finland R&D Department, Novia University of Applied Sciences, Vaasa, Finland d Jerzy Haber Institute of Catalysis and Surface Chemistry, Polish Academy of Sciences, Kraków, Poland e Department of Geology and Mineralogy, Åbo Akademi University, Åbo, Finland b c

H I G H L I G H T S • • • •

Acid sulfate soils cause environmental damage by releasing acid and metals. Acid sulfate soils treated with CaCO3 and Ca(OH)2 released solutions with higher pH. The microbial populations associated with treated and untreated soils were similar. The data can be used to design mitigation strategies to treat acid sulfate soils.

a r t i c l e

i n f o

Article history: Received 11 January 2015 Received in revised form 13 April 2015 Accepted 13 April 2015 Available online 28 April 2015 Editor: F.M. Tack Keywords: Acid sulfate soil Acid Metal Molecular phylogeny Acidophile

a b s t r a c t Potential acid sulfate soils contain reduced iron sulfides that if oxidized, can cause significant environmental damage by releasing large amounts of acid and metals. This study examines metal and acid release as well as the microbial community capable of catalyzing metal sulfide oxidation after treating acid sulfate soil with calcium carbonate (CaCO3) or calcium hydroxide (Ca(OH)2). Leaching tests of acid sulfate soil samples were carried out in the laboratory. The pH of the leachate during the initial flushing with water lay between 3.8 and 4.4 suggesting that the jarosite/schwertmannite equilibrium controls the solution chemistry. However, the pH increased to circa 6 after treatment with CaCO3 suspension and circa 12 after introducing Ca(OH)2 solution. 16S rRNA gene sequences amplified from community DNA extracted from the untreated and both CaCO3 and Ca(OH)2 treated acid sulfate soils were most similar to bacteria (69.1% to 85.7%) and archaea (95.4% to 100%) previously identified from acid and metal contaminated environments. These species included a Thiomonas cuprina-like and an Acidocella-like bacteria as well as a Ferroplasma acidiphilum-like archeon. Although the CaCO3 and Ca(OH)2 treatments did not decrease the proportion of microorganisms capable of accelerating acid and metal release, the chemical effects of the treatments suggested their reduced activity. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Sulfide rich sediments are widespread in many low-lying coastal areas worldwide and are especially prevalent in Australia, Asia, and Europe (Andriesse and van Mensvoort, 2006). These sediments are referred to as potential acid sulfate soil (PASS) material when they

Abbreviations: PASS, potential acid sulfate soil; ASS, acid sulfate soil; RFLP, restriction fragment length polymorphism. ⁎ Corresponding author. E-mail addresses: [email protected] (X. Wu), pekka.sten@vamk.fi (P. Sten), sten.engblom@novia.fi (S. Engblom), [email protected] (P. Nowak), posterho@abo.fi (P. Österholm), [email protected] (M. Dopson).

http://dx.doi.org/10.1016/j.scitotenv.2015.04.049 0048-9697/© 2015 Elsevier B.V. All rights reserved.

have not been oxidized (Rabenhorst et al., 2006). Natural phenomena (e.g., land uplift) or artificial drainage (e.g., for agricultural use) can expose the sediments to atmospheric oxygen and result in metal sulfide oxidation and the subsequent formation of acid sulfate soil (ASS) (Sundström et al., 2002). These soils typically have a pH of 2.5–4.0 (Åström, 2001). Finland has approximately 3000 km2 of ASS in costal farmland developed from sulfide rich sediments (Beucher et al., 2013) that constitutes the largest ASS area in Europe (Andriesse and van Mensvoort, 2006). As a result of the oxidation of iron sulfides, large amounts of acid and high level of metals (e.g., Al, Cd, Ni, and Mn) are released (Åström, 2001; Boman et al., 2010; Nordmyr et al., 2008; Sundström et al., 2002) that leads to severe environmental damage. For instance, massive acidic drainage from ASS has caused large fish kills in western coastal areas of Finland as recently as 2006.

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Sulfur- and iron-oxidizing microorganisms are widely found in many acidic environments such as acid mine drainage (Hallberg, 2010), sulfide mine tailings (Rzhepishevska et al., 2005), and ASS (Wu et al., 2013). The role of extreme acidophiles in catalyzing the iron and/or inorganic sulfur compound oxidation leading to the release of acid and metals from metal sulfides is well established (Vera et al., 2013). These microorganisms need to be able to survive in the low pH and high metal concentrations generated by sulfide mineral dissolution and their pH homeostasis (Slonczewski et al., 2009) and metal resistance (Dopson et al., 2003, 2014; Dopson and Holmes, 2014) strategies have been reviewed. A microbiological study of ASS in boreal environments identified several 16S rRNA gene sequences similar to known acidophilic iron and inorganic sulfur oxidizing microorganisms, suggesting that the acid and metal releases are accelerated by these microorganisms (Wu et al., 2013). In addition, 16S rRNA gene sequence analysis suggested the presence of acidophiles adapted to the low temperatures found in Finland. These included the iron- and sulfur-oxidizing species Acidithiobacillus ferrivorans and 16S rRNA gene sequences related to clones detected in Arctic soil where rocks had been weathered by chemolithoautotrophic microorganisms (Borin et al., 2010). A second study of boreal ASS identified the microbial populations, including methanotrophs, in ASS and non-ASS and demonstrated the capacity for greenhouse gas emission if PASSs are drained (Šimek et al., 2014). Due to the significant damage caused by acid and metal release from acid mine drainage, various remediation techniques have been developed including chemical treatment with lime, calcium carbonate, sodium hydroxide, etc., to neutralize the total acidity of the acid mine drainage liquors (Johnson and Hallberg, 2005). Several strategies have also been studied to prevent pyrite oxidation or to remediate the acidic groundwater caused by ASS (Golab et al., 2006; Indraratna et al., 2005). Techniques such as controlled drainage, surface liming, and lime-filter drainage have also been tested on ASS in Finland (Åström et al., 2007). However, the applied methods only appear to have had a minor effect on the release of acid and metals. In this study, a new strategy of sub-surface calcium carbonate and calcium hydroxide treatment on metal and acid generation at the Risöfladan experimental field, Finland was evaluated in laboratoryscale column experiments. Compared to previous studies, this approach specifically targeted the hydrologically active macropores in the critical soil horizon. In addition, the impact of the chemical treatments on the microbial community present in the ASS was evaluated. The results can be utilized for the long term mitigation of environmental damage caused by boreal ASS.

The sample (from the soil horizon extending from 70 to 85 cm below the surface) was inserted into a cylindrical rubber membrane that was pressed against the soil during the laboratory experiment by submerging in water and applying an external pressure to prevent bypass flow between the membrane and the soil sample. Even though this soil horizon was designated as oxidized, it still contained significant amounts of pyrite (Boman et al., 2010). 2.2. Chemical methods and instrumentation The experimental set-up for the column tests is shown in Fig. 1. The two chemical (CaCO3 and Ca(OH)2) treatments were carried out in duplicate on separate soil samples. For each of the treatments, one representative experiment is presented and the data from the replicate experiments are in the Supplemental files. The experiment began with passing ~ 10 dm3 pure water through the column. During the experiments, pH, temperature-corrected (to 25 °C) conductivity, and oxidation-reduction potential (ORP) versus an Ag/AgCl (3.5 M KCl) reference electrode were continuously registered in situ at the outlet using an YSI Professional Plus instrument (YSI Inc., Yellow Springs, OH 45387, U.S.A.) with a flow-through cell (Quatro Cable Flow Cell) from the same manufacturer. The recorded ORP values were recalculated to potentials (EH) versus the SHE by the addition of 205 mV, before being converted to pe values (pe = EHF / (2.303RT)). The column effluent was collected in a cylinder open to the atmosphere. Every few hours the volume of the outlet solution was measured and ex situ measurements of pH (SchottGeräte CG822 or Radiometer PHM210) and conductivity (Radiometer 2− CDM210) were performed. The concentrations of Cl−, NO− 3 , and SO4 were measured by ion chromatography (Dionex ICS 1100) and Al by UV spectrophotometry (Hach Lange DR 3900). Several soil samples were subjected to X-ray diffraction analysis using the PANalytical X'Pert Pro system with monochromatic Cu Kα radiation. The phase composition of the samples was identified using the International Centre for Diffraction Data database. All experiments and measurements were performed at ambient temperatures (20–21 °C). Concentrations of selected metals were determined by Inductively Coupled Plasma Mass Spectrometry or by Inductively Coupled Plasma/Optical Emission Spectrometry. The treatment chemicals, CaCO3 (trade name Nordkalk C2,

2. Methods 2.1. Study site and sampling ASS samples for chemical treatments were collected at the Risöfladan experimental field in Vaasa, Finland (63.05°N, 21.71°E) by removing the plow layer (0–30 cm) and then pushing polyethylene tubes into the ground with an excavator. Details of the study area (Åström et al., 2007; Boman et al., 2010) and the physicochemical parameters of the Risöfladan experimental field (Wu et al., 2013) have been previously reported. In brief, the soil below the plow layer consisted of clay with a well-developed structure with an abundance of iron oxide coatings. The sulfur content was in the order of ≥ 0.2% (wt/vol) and the pH was ~4 in all soil samples due to significant sulfide oxidation (Wu et al., 2013). In total, 24 cores were used in preliminary studies, method development, and the five cores for the experiments reported here. A new core was utilized for each experiment by slicing a soil sample from the appropriate soil horizon as described below. The tubes with the soil profiles were immediately sealed at both ends to prevent the ingress of air and delivered to the laboratory. A cylinder-shaped sample was obtained by cutting the 14.2 cm inner-diameter polyethylene tube into a 15 cm long piece and carefully removing the soil core.

Fig. 1. Experimental set-up for the leaching tests. The maximum hydrostatic pressure employed corresponded to a water column of circa 3.7 m.

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ground calcium carbonate with median particle diameter d50 = 2.5 μm) and Ca(OH)2 (trade name Nordkalk SL 90T, slaked lime with a Ca(OH)2 content of N 93% and with 96.3% of particle diameters b 90 μm) were of technical grade and supplied by Nordkalk Corporation (www.nordkalk. com). Analyses of suspensions/solutions of the treatment chemicals showed that they did not significantly contribute to the metal and sulfur concentrations monitored in the experiments. High-purity water, previously equilibrated with the atmosphere, was used in the leaching experiments and for preparing the suspensions and solutions. 2.3. Chemical treatments Column treatment experiments were carried out by the application of either CaCO3 or Ca(OH)2 suspensions/solutions to the ASS as a neutralizing agent. Preliminary experiments with 1, 5, and 10 g dm− 3 CaCO3 suspensions (Sten et al., 2012) showed that the 1 g dm−3 suspension exerted a very little influence on pH and composition of the outlet solution but 5 g dm−3 and 10 g dm−3 suspension led to similar results. Therefore, 10 g dm−3 suspension was used in further experiments. The inlet CaCO3 suspension (10 g dm−3) had an initial pH of 9.6 and a conductivity of 51 μS cm−1. The saturated Ca(OH)2 solution had an initial pH of 12.6 and a conductivity of 7.4 mS cm−1. The soil samples were leached with water before and after CaCO3 or Ca(OH)2 suspensions/solutions were passed through the soil samples under hydrostatic pressure. Due to differences in permeability between soil samples, the rates of solution that flow through the columns were dissimilar and therefore, the volumes of water as well as CaCO3 and Ca(OH)2 suspension/solution varied between experiments. The amount of water that passed through the soil column prior to the chemical treatment was approximately equal to 1 to 2 years of percolation in the field. Although the amount of water that passed through the column was equivalent to 1 to 2 years, it occurred over a much shorter time frame. This amount of time was probably insufficient for all chemical reactions that would have significantly increased the pool of acidity and metals to have taken place. However, the results suggested that it was sufficient for an amount of acid and metal release that typically occurs during one year. However, it would likely have been sufficient to affect acidophilic microorganisms as most species are unable to survive at neutral pH. 2.4. pe–pH diagrams pe–pH diagrams were drawn using PhreePlot (http://www. phreeplot.org/), containing an embedded version 3 of the PHREEQC program (Parkhurst and Appelo, 1999). Equilibrium constants used were those provided in the wateq4f database (part of the PhreePlot/PHREEQC software; the wateq4f.dat file used is dated 21 August 2012), except for K-jarosite where the log K value of −11.0 (KFe3(SO4)2(OH)6 + 6 H+ → K+ + 3Fe3 + + 2SO24 − + 6H2O) was used (Baron and Palmer, 1996) and schwertmannite where the log K value of 20 was used (Fe8O8(OH)6SO4 + 22 H+ → 8Fe3+ + SO2− 4 + 14H2O). It is particularly difficult to find agreement in the literature for the metastable schwertmannite log K value. However, the value used in this work lies in the interval (18.0 ± 2.5) given by Bigham et al. (1996). 2.5. Molecular phylogenetic analysis The molecular phylogeny of microbial populations from both the untreated and treated soils was analyzed as previously described (Wu et al., 2013) except that the DNA was directly extracted using the PowerSoil DNA isolation kit (MOBIO) according to the manufacturer's instructions. Microbial community DNA was extracted from a thin layer of soil scraped from the surface of cracks within the soil through which the majority of the CaCO3 and Ca(OH)2 suspensions flowed. In addition, as a control to evaluate if the microbial community was different between treated and untreated areas of soil, community DNA was extracted from an area without cracks through which little or no

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CaCO3 or Ca(OH)2 suspensions/solutions had flowed. Partial 16S rRNA gene sequences were amplified from the community DNA samples with both archeal and bacterial primers, cloned into the pGEM-T Easy Vector System (Promega), and transformed into Escherichia coli (Wu et al., 2013). Plasmid preparations from the transformed E. coli were PCR amplified and individual phylotypes identified by restriction fragment length polymorphism (RFLP). The representative clones were sequenced by Macrogen. The sequences obtained were checked using DECIPHER to remove the chimeric sequences (Wright et al., 2012), edited in Geneious version 6.0.6 (Biomatters Ltd Auckland, New Zealand), and compared with the GenBank database using BLAST before constructing the phylogenetic trees with Molecular Evolutionary Genetic Analysis version 5.1 (Tamura et al., 2011). 16S rRNA gene sequences were submitted to GenBank with the accession numbers KJ768441– KJ768609. 3. Results and discussion 3.1. Soil physicochemical characteristics during washing with water The flow of water through the soil column was highly irregular both in space and in time and preliminary experiments using Alizarin Red S dye showed that water only flowed through the channels and fissures (Sten et al., 2012). Similarly, the CaCO3 and Ca(OH)2 were not evenly distributed in the whole/bulk sample, but rather the majority of the neutralizing agent passed through and/or remained in the hydrologically active macropores. As the reactions producing acid and liberating metals predominantly take place at these sites, the soil material most likely to be oxidized was predominantly treated. Despite a constant hydrostatic pressure the flow through the soil columns steadily decreased, but in some rare cases an increase in the flow rate was observed. The decrease in the flow rate was probably caused by clogging of the channels by particles flowing with the water while the increase could be due to opening of new channels. The initial conductivity decreased sharply and reached a value between 150 and 300 μS cm− 1 during the initial flushing with 5 to 14 dm3 of water. In a separate control experiment, a prolonged flushing with a total of 40 dm3 water further reduced the conductivity to 80 μS cm−1 (Fig. 2A). The decrease in conductivity was also seen during the initial stage of the replicate CaCO3 and Ca(OH)2 treatments (number of replicates (n) = 4), and was mainly due to flushing out of the sulfate, chloride, and nitrate anions, as well as the base cations. After having passed a sufficiently large volume of water for the nitrate and chloride ion concentrations to decrease practically to zero, the concentration of sulfate stabilized at circa 90 mg dm−3. This suggested that the sulfate concentration was stabilized by dissolution equilibrium. The similar behavior of the conductivity and sulfate concentration (Fig. 2A) suggested that the main anionic species in solution was sulfate, although the K+ ion concentration also stabilized at a relatively high concentration (Fig. 2B). The decrease in conductivity was mainly due to flushing out of sulfate ions from the oxidation of iron sulfides present in the soil, chloride ions from when the soil was a sea bottom, nitrate ions from fertilizers, and calcium and magnesium ions from the dissolution of trace calcareous minerals. Despite the instability of the flow and the decrease in conductivity caused by flushing of ionic species from the soil, the pH of the outlet solution during the initial flushing period was largely constant (pH 3.8 to 4.4; n = 5) both within and between experiments. The most probable explanation was due to a dissolution/precipitation equilibrium between jarosite and schwertmannite (Bigham et al., 1996; Collins et al., 2010; España, 2007; Marescotti et al., 2012; Regenspurg et al., 2004) as described by Eq. (1).

þ

2−

þ

8KFe3 ðSO4 Þ2 ðOHÞ6 →3Fe8 O8 ðOHÞ6 SO4 þ 8K þ 13SO4 þ 18H þ 6H2 O

ð1Þ

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the ratio of sulfate to potassium in the effluent of the column was always much higher than that predicted by Eq. (1). This suggested that in addition to jarosite other sulfate-containing minerals were also present in the soil. Considering the relatively high concentration of calcium ions (Fig. 2B) the most probable mineral was gypsum. Finally, some free sulfuric acid might be formed as a result of iron sulfide oxidation, the concentration of which depends on the amount of sulfides in the soil and the availability of oxygen. The laboratory leaching experiments may be compared to data collected from a neighboring field experiment. In the nearby field the runoff was 244–278 mm a−1 (circa 40% of precipitation) (Österholm et al., in press), corresponding to a percolation of circa 4000 mL a−1 in the 14.2 cm diameter soil column. Leached amounts of metals per hectare and year were estimated using average concentrations in the outlet solution during the initial flushing stage (prior to the chemical treatment) of soil columns. Concentrations of calcium were relatively high, on average 25 mg dm−3 corresponding to an annual leaching of 60 kg per hectare. High values were also observed for potassium (6 mg dm−3/ 16 kg ha−1 a−1), aluminum (4 mg dm−3/10 kg ha−1 a−1), and manganese (1.3 mg dm− 3/3 kg ha− 1 a−1), while the concentration and leaching of iron was relatively low (b 1 mg dm−3/b 1 kg ha−1 a−1). Considering that the ASS soil profile in the field was about tenfold longer

Fig. 2. Change of ex situ (e.s.) pH, ex situ (e.s.) conductivity (κ) and the concentration of sulfate, nitrate, and chloride ions (A) as well as selected metal ion concentrations (B) during flushing of the soil sample with water. “×” marks the concentration at the limit of quantification. The total duration of the experiment was 18 days, leading to an average flow rate of 94 mL h−1.

Indeed, jarosite (along with quartz and a less certain identification of a minor component of albite) was observed in samples from the Risöfladan experimental field by X-ray diffraction (Supplemental File 1) and has previously been reported from similar ASS (Joukainen and Yli-Halla, 2003; Nordmyr et al., 2006). The other side of the schwertmannite stability region borders either goethite or ferrihydrite (Eqs. (2) & (3)). 2−

Fe8 O8 ðOHÞ6 SO4 þ 2H2 O→8FeOOH þ SO4

2−

Fe8 O8 ðOHÞ6 SO4 þ 10H2 O→8FeðOHÞ3 þ SO4

þ

þ 2H

ð2Þ þ

þ 2H

ð3Þ

According to Bigham et al. (1996), the pH of the jarosite/ schwertmannite equilibrium lies (depending on assumed solubility product) between circa pH 1.8 and 3 and the schwertmannite/ ferrihydrite equilibrium pH lies between circa pH 4.4 and 5.7 while Schmiermund (2008) states that the stability region of K-jarosite extends to circa pH 5. Even though the comparison should be used with caution (the dynamic situation of an experiment versus the equilibrium assumed in a pe–pH diagram) and taking into account that the border lines between jarosite/schwertmannite and schwertmannite/ferrihydrite are highly dependent on the logK values used, it was evident that experimental pH and redox conditions rendered high importance to the jarosite and schwertmannite phases (Fig. 3A & B). Further evidence supporting the jarosite/schwertmannite dissolution/precipitation equilibrium buffering the pH was that large amounts of sulfate and potassium were liberated by the CaCO3 and Ca(OH)2 treatments (described in Section 3.2). However, it must be noted that

Fig. 3. pe–pH diagram during initial flushing with water (closed circles) in a soil with no calcite added (A) and after calcite treatment (initial points are circled), and the final water treatment (final points are circled) (B). The duration of the experiment was 22 days.

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(circa 1.5 m), these results were in line with the Nordmyr et al. (2008) who estimated the annual metal flux from a nearby field. This indicated that the column experiments reflected leaching in field conditions relatively well.

3.2. Soil behavior during the CaCO3 and Ca(OH)2 treatments When water was replaced by the calcium carbonate suspension the conductivity and pH increased (Fig. 4A). The increase in conductivity was accompanied by a rise in the sulfate (Fig. 4A) and potassium ion (Fig. 4B) concentrations which may be ascribed to the enhanced dissolution of jarosite (Eq. (4)). þ



2KFe3 ðSO4 Þ2 ðOHÞ6 þ 3CaCO3 þ 3H2 O→2K þ 3Ca

2−

þ 4SO4 þ 6FeðOHÞ3 þ 3CO2

ð4Þ

During soil treatment with the CaCO3 suspension almost all measured pe–pH points lay in the range of the stability of schwertmannite (Fig. 3B). Carbon dioxide is liberated in reaction (4), which leads to over-saturation of the solution with CO2 and possible blocking of pores by CO2 bubbles. This gas was liberated at the outlet of the column which caused an increased pH. Indeed, the ex situ pH was always higher than the in situ pH during flushing with CaCO3 suspension (Fig. 4A), showing a clear time dependency and approaching pH 8. An increase in manganese ion concentration was also observed (Fig. 4B), which may be ascribed to the susceptibility of the manganese ions to form complexes with carbonate ions. The decrease in aluminum ion concentration (Fig. 4B) was beneficial as these ions are very toxic for aquatic fauna. It was also notable that concentrations of nitrate and chloride

Fig. 4. Change of in situ (i.s.) pH and ex situ (e.s.) pH; in situ (i.s.) and ex situ (e.s.) conductivity (κ) and the concentration of sulfate, nitrate, and chloride ions (A) and selected metal ion concentrations (B) during flushing of the soil sample with water and with the CaCO3 suspension. “×” marks the concentration at the limit of quantification. The total duration of the experiment was 19 days. The flow rate during the initial flushing was approximately 70 mL h−1, during treatment 30 mL h−1, and during final flushing the flow rate increased again and approached the rate during the initial flushing.

219

anions increased upon the introduction of the CaCO3 suspension, probably due to anion exchange with CO2− 3 and possibly to some extent due to a decreased anion exchange capacity caused by the higher pH (Fig. 4A). The evolution of carbon dioxide inside the soil sample may also explain the instability in the flow rate and the fluctuations in pH observed after the introduction of the CaCO3 suspension followed by washing with water. When the calcium carbonate suspension was again replaced by water, the pH was maintained at a high value, despite that the conductivity decreased to slightly higher than during washing the sample with water for a comparable period of time (Figs. 2A & 4A). A repeat experiment for treatment with CaCO3 suspension gave similar results and is described in Supplemental File 2. The results for duplicate treatments with saturated Ca(OH)2 solution were similar to washing with CaCO3 suspension (Fig. 5 & Supplemental File 3). The differences between the two treatments were that the pH stabilized at a much higher value during washing with saturated Ca(OH)2 and no CO2 evolution was expected (Eq. (5)). þ



2KFe3 ðSO4 Þ2 ðOHÞ6 þ 3CaðOHÞ2 →2K þ 3Ca

2−

þ 4SO4

þ 6FeðOHÞ3

ð5Þ

In addition, in contrast to during CaCO3 treatment the concentration of aluminum temporarily increased with the introduction of Ca(OH)2 potentially due to the formation of Al(OH)− 4 at high pH (Fig. 6). However, subsequent washing with water decreased the pH and consequently the aluminum concentration. 3.3. Molecular phylogeny of the untreated and treated soils Archeal and bacterial 16S rRNA gene sequences were amplified from the duplicate untreated and treated soils for each treatment

Fig. 5. Change of ex situ (e.s.) pH, ex situ (e.s.) conductivity (κ) and the concentration of sulfate, nitrate, and chloride ions (A) and selected metal ion concentrations (B) during flushing of the soil sample with water and with the solution of Ca(OH)2. The total duration of the experiment was 18 days. The flow rate was approximately 50 mL h−1 but decreasing during the last 2 L.

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Fig. 6. Change in Al concentration during flushing of the soil sample with water, CaCO3 suspension, and with the saturated solution of Ca(OH)2.

(Supplemental File 4). The untreated and CaCO3 treated soils had a very high ratio of archeal and bacterial clones most similar to sequences previously identified from acidic or metal rich related environments (Supplemental Files 5 to 8). These percentages of acidic or metal rich related clones to total clones were 100 ± 0% and 96.4 ± 5.1%, respectively for the archaea and 77.7 ± 26.7% and 85.7 ± 12.4%, respectively for the bacterial clones (Table 1). Similar data were also observed for the Ca(OH)2 experiment with 100 ± 0% archeal clones previously associated with acidic and metal rich milieu in the untreated soil compared with 95.4 ± 1.0% in the treated soil and 84.6 ± 8.9% and 69.1 ± 11.8% bacterial related sequences in the Ca(OH)2 treatment, respectively (Table 1). The high relative number of identified 16S rRNA gene clone sequences most similar to previously identified acidic and metal related microorganisms suggests that microbes capable of catalyzing acid and metal release were still present in ASS after treatment. This was likely due to these samples being sampled from treated fissure surfaces where oxidation had already occurred, lowering the pH, and allowing acidophilic microorganisms to proliferate. The CaCO3 and Ca(OH)2 treatments decreased the effluent acidity but not the ratio of 16S rRNA gene clone sequences related to previously identified acidophilic microorganisms. However, they may have inhibited the activity of the acidophilic microorganisms to catalyze pyrite oxidation by raising the pH above the optima for acidophilic microorganisms (Dopson, 2012). 16S rRNA gene sequences most similar to uncultured Crenarchaeote clones contributed a large proportion of the archaea identified from the community, including several archeal clones in both untreated and treated soils that were similar to known acidophiles and clones previously isolated from acidic, metal- and sulfur-containing environments

Table 1 16S rRNA gene clones identified in the treated fissure surfaces and untreated regions of the soil columns for the duplicate experiments in both treatments. Total clones in duplicate experiments

Acid environment related 16S rRNA genes

% of total acidic or metal rich related clones

CaCO3 treatment Archaea Untreated Treated Bacteria Untreated Treated

33 14 17 13

29 28 29 36

31 ± 2.8 20.5 ± 10.6 19 ± 12.7 22 ± 17

100 ± 0 96.4 ± 5.1 77.7 ± 26.7 85.7 ± 12.4

Ca(OH)2 treatment Archaea Untreated Treated Bacteria Untreated Treated

37 19 37 31

31 26 44 28

34 ± 4.2 21.5 ± 4.9 34.5 ± 7.8 20.5 ± 4.9

100 ± 0 95.4 ± 1.0 84.6 ± 8.9 69.1 ± 11.8

(Supplemental Files 5 & 7). These included 34U_A33, 34T_A13, 25U_A23, 25U_A45, 25T_A14 and 25T_A33 which were all the most similar (94% to 99%) to Ferroplasma acidiphilum, which is an ironoxidizing, acidophilic archaea (Dopson et al., 2004) and 25T_A35 that was 88% similar to an uncultured Thermoplasma sp. identified from an acidic forest soil (Kemnitz et al., 2007). Several bacterial clones in both treatments had high similarity to clones identified in our previous study of the Risöfladan experimental field (Supplemental Files 6 & 8). These included (but were not limited to) 35T_B67, 34T_B11, 34T_B14, 35U_B8, and 34U_B6 most similar to V2C2 cloned from the acidic oxidized layer; 25T_B29 most similar to clone C4A8 from an enrichment culture on yeast extract; and 34T_B28 most similar to clone B4H3 from an acidic tetrathionate enrichment culture (Wu et al., 2013). Additional bacterial 16S rRNA gene sequences identified in this study related to acidophilic microorganisms and other cold environments included 34U_B4 that was 99% similar to an Acidobacteria clone (de Castro et al., 2013); 35U_B35 was 98% similar to Thiomonas cuprina, a ferrous iron and arsenite oxidizer (Kelly et al., 2007); 35U_B9 was 98% similar to Acidocella sp. (Kay et al., 2013); and 34U_B36, 35U_B14, 34T_B18, 26U_B68, 25U_B13, and 26U_B58 most similar to a macroscopic streamer bacterium isolated from cold, acidic acid mine drainage (Hallberg et al., 2006). 4. Conclusions Initial flushing of the ASS samples resulted in leachate pH values of circa 4, indicating an existing solution chemistry dominated by the jarosite/schwertmannite equilibrium. Both the CaCO3 and Ca(OH)2 treatments raised the pH of the water passing through the soil columns, mitigating both acid discharge and metal leaching from the ASS samples. Even after the treatment solutions/suspensions were replaced by water, leachate pH values remained at levels considerably higher than during the initial flushing stage. These results indicated that the treatments tested could be successful as part of a large-scale mitigation strategy. However, the treatments did not decrease the number of clones sequenced from acidophilic microorganisms. The clones were taken from fissure surfaces that had already been oxidized and would have had a low pH. The subsequent increase in the pH would likely inhibit microbial activity and potentially killed the acidophilic microorganisms. Further studies are required to investigate how the activity of the acidophilic microorganisms changes during the treatment. This would verify if the decrease in acid release as a result of the mitigation strategies was due to purely chemical effects or if they also inhibited biological catalysis of metal sulfide dissolution. Acknowledgments This work was part of the PRECIKEM (Chemical precision treatment of acid sulfate soils to prevent acid formation) project. The project was funded by the European Agricultural Fund for Rural Development via the Rural Development Programme for Mainland Finland 2007–2013. This programme was administrated by the Centre for Economic Development, Transport and the Environment in Ostrobothnia (grant number 5486/3560-2010). Co-funding was provided by the Field Drainage Association, K.H. Renlunds stiftelse, Maa- ja vesitekniikan tuki, Central Union of Agricultural Producers and Forest Owners (both by the Foundation of this organization as well as the regional chapter in South Ostrobothnia), and Österbottens Svenska Producentförbund. Additional funding for expert assistance, equipment, etc., was provided by the Oiva Kuusisto Säätiö, K.H. Renlunds stiftelse, Aktiastiftelsen i Vasa, and Stiftelsen Handlanden Gustaf Svanljungs Donationsfond. Aid and support was also provided by the Nordkalk company. The funding sources had no role in experimental design or interpretation of the data. We would also like to express our gratitude to Professor Mats Åström, Environment Councilor Pertti Sevola and Senior Officer Karl-Erik Storberg for

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Impact of mitigation strategies on acid sulfate soil chemistry and microbial community.

Potential acid sulfate soils contain reduced iron sulfides that if oxidized, can cause significant environmental damage by releasing large amounts of ...
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