989

AbstractFollowing the break of a severe drought in the Murray–Darling Basin, rising water levels restored subaqueous conditions to dried inland acid sulfate soils with sulfuric horizons (pH <3.5). Equilibrium dialysis membrane samplers were used to investigate in situ changes to soil acidity and abundance of metals and metalloids following the first 24 mo of restored subaqueous conditions. The rewetted sulfuric horizons remained severely acidified (pH ~4) or had retained acidity with jarosite visibly present after 5 mo of continuous subaqueous conditions. A further 19 mo of subaqueous conditions resulted in only small additional increases in pH (~0.5–1 pH units), with the largest increases occurring within the uppermost 10 cm of the soil profile. Substantial decreases in concentrations of some metal(loid)s were observed with time most likely owing to lower solubility and sorption as a consequence of the increase in pH. In deeper parts of the profiles, porewater remained strongly buffered at low pH values (pH <4.5) and experienced little progression toward anoxic circumneutral pH conditions over the 24 mo of subaqueous conditions. It is proposed that low pH conditions inhibited the activity of SO4

2−–reducing bacteria and, in turn, the in situ generation of alkalinity through pyrite production. The limited supply of alkalinity in freshwater systems and the initial highly buffered low pH conditions were also thought to be slowing recovery. The timescales involved for a sulfuric horizon rewetted by a freshwater body to recover from acidic conditions could therefore be in the order of several years.

Porewater Geochemistry of Inland Acid Sulfate Soils with Sulfuric Horizons Following Postdrought Reflooding with Freshwater

Nathan L. Creeper, Paul Shand, Warren Hicks, and Rob W. Fitzpatrick*

This study examines the changes to soil porewater chemistry that occurred when previously desiccated active acid sulfate soils (ASS) with a sulfuric horizon

(pH <3.5) (Soil Survey Staff, 2014) were rapidly submerged by a large freshwater body. Acidity, metals, and metalloids (Al, As, Fe, Mn, Ni, Zn) made available from the oxidation of pyrite contained in potential ASS with sulfidic material (pH >3.5) or subsequent processes, such as the acid dissolution of layer silicate clays, can be mobilized during rewetting and potentially result in damage to surrounding ecosystems (Astrom and Astrom, 1997; Dent, 1986; Macdonald et al., 2004; Nyberg et al., 2012; Nystrand and Osterholm, 2013). The most significant risks of environmental degradation usually occur following the rewet-ting of desiccated and acidified (pH <3.5) active ASS but before the reestablishment of reducing conditions where dilution and neutralization toward circumneutral pH can immobilize some metal species and encourage the reformation of sulfide miner-als. The potential to suppress the mobilization of trace metals and encourage the reformation of iron sulfide minerals is why the rewetting of soils with sulfuric horizons has been suggested as a low-cost, landscape-scale remediation technique by sev-eral workers in Australian coastal environments (Burton et al., 2008; Burton et al., 2005; Johnston et al., 2009b, 2009c; Keene et al., 2010, 2011; Simpson et al., 2010; Virtanen et al., 2014). Although the desired outcomes and fundamental processes of the technique remain the same, the rewetting of sulfuric hori-zons in freshwater systems is likely to have significant differences to the same approach when used in coastal systems ( Johnston et al., 2010a, 2012, 2010b, 2011; Portnoy and Giblin, 1997a, 1997b) where, for example, the tidal cycle provides a constant diurnal resupply of alkalinity. This study provides observations specific to the behavior of active inland ASS (IASS) inundated by a freshwater body. There are few publications that are directly applicable to the freshwater environment of this study (Hicks et al., 2003, 2009b; Mosley et al., 2014a, 2014b; Shand et al., 2010; Virtanen et al., 2013). Investigation into the behavior of rewet-ted IASS with sulfuric horizons is important so that these types

Abbreviations: AHD, Australian Height Datum; ASS, acid sulfate soil; Eh, redox potential; IASS, inland acid sulfate soil; MDB, Murray–Darling Basin; SWI, soil–water interface.

N.L. Creeper, P. Shand, and R.W. Fitzpatrick, Acid Sulfate Soils Centre, EES, the Univ. of Adelaide, Private Bag No 1, Glen Osmond, South Australia, Australia, 5064; N.L. Creeper, P. Shand, W. Hicks, and R.W. Fitzpatrick, CSIRO Land and Water, Private Bag No 2, Glen Osmond, South Australia, Australia, 5064; P. Shand, School of the Environment, Flinders Univ., PO Box 2100, Adelaide, Australia, 5001. Assigned to Associate Editor Christian Stamm.

Copyright © American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America. 5585 Guilford Rd., Madison, WI 53711 USA. All rights reserved. No part of this periodical may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. J. Environ. Qual. 44:989–1000 (2015) doi:10.2134/jeq2014.09.0372 Supplemental material is available online for this article. Received 3 Sept. 2014. Accepted 18 Dec. 2014. *Corresponding author (rob.fitzpatrick@csiro.au).

Journal of Environmental QualityWETLANdS ANd AQuATIC PROCESSES

TECHNICAL REPORTS

990 Journal of Environmental Quality

of scenarios can be effectively managed, minimizing the poten-tial of any damage to surrounding ecosystems.

Pyrite-bearing soils, known as potential ASS, can be found in the inland wetlands, streams, rivers, and lakes of Australia, including those systems associated with the Murray–Darling Basin (MDB) (Creeper et al., 2013; Earth Systems, 2010; Fitzpatrick et al., 2009c; Glover et al., 2011; Hall et al., 2006; Lamontagne et al., 2006; Wallace et al., 2006; Welch et al., 2006). The study area is located in the lower reaches of the Finniss River and Currency Creek catchments adjacent to Lake Alexandrina and on the southwestern edge of the MDB in South Australia. The catchments drain the Eastern Mount Lofty ranges and flow into the Goolwa Channel near the mouth of the Murray River (Fig. 1).

There have been a number of water regime changes that ultimately led to the formation of sulfidic materials and subsequent sulfuric horizons in the IASS in this study (Fitzpatrick et al., 2012). The following is a summary of the water regime changes and their effects relating to ASS characteristics. In the 1920s, locks and barrages were installed along the Murray River to regulate the water level and facilitate irrigation and navigation along its length. The regulation of high water levels replaced the natural wetting and drying regime with prolonged subaqueous conditions throughout large areas of the MDB. The almost permanent subaqueous conditions that lasted from ca. 1920s to 2003, an 80-yr period, resulted in the accumulation of sulfides above usual levels in the subaqueous and marginal soils of the wetlands, lakes, and rivers of the MDB.

In 2003, a severe drought affecting much of the MDB, combined with over allocation of water resources, led to low water levels in the Lower Lakes region including the Finniss River and Currency Creek tributaries. Water levels were at their lowest from 2007 to 2009 and large areas of previously submerged soils were exposed to the atmosphere. Sulfides that had accumulated in these soils during the previous 80 yr of continuous subaqueous conditions were then oxidized, resulting in the formation of approximately 2000 ha of soils with sulfuric horizons in the lower reaches of the Finniss River and Currency Creek catchments (Fitzpatrick et al., 2009a).

Due to the formation of large quantities of ASS with sulfuric horizons and the expected continued lowering of water levels, a water regulator was installed in August 2009 to prevent further areas of subaqueous potential IASS being exposed in the Finniss River and Currency Creek catchments (Clayton regulator, Fig. 1). Shortly thereafter, the pumping of water from Lake Alexandrina into the tributaries, combined with large unseasonal rainfall events, resulted in the rewetting of the exposed IASS with sulfuric horizons in the Finniss River and Currency Creek in September 2009. Water levels stabilized at ~0.5 m Australian Height Datum (mean sea level assigned value of 0 m AHD) in November 2009 and remained between 0.5 m AHD and 0.9 m AHD through to the conclusion of this work.

Materials and MethodsSelection of Sampling Locations

A total of four sampling locations were selected to study chemical changes following the rewetting of sulfuric horizons: two sampling locations in the Finniss River, Fin_N (−35.407

lat., 138.832 long.) and Fin_S (−35.449 lat., 138.863 long.); and two sampling locations in Currency Creek, Cur_N (−35.459 lat., 138.778 long.) and Cur_S (−35.481 lat., 138.853 long.) (Fig. 1). The four sampling locations discussed in this paper coincide with previous research and are a subset of 39 geographically distributed and locally representative soil profiles (Fitzpatrick et al., 2009a). Sample labeling was simplified for this paper and differs from the labeling used in previous studies. Previous labels corresponding to those used in this study are provided in Supplemental Table S1. The 39 soil profiles were originally surveyed in November 2008 to determine the properties and extent of ASS in the lower reaches of Finniss River and Currency Creek catchments and resurveyed approximately 1 yr later (Fitzpatrick et al., 2009a, 2009b, 2011). The profiles at the four sampling locations discussed in this paper were selected because they were active IASS with a sulfuric horizon and represented both dominantly clay-textured (Fin_N and Fin_S) and dominantly sand-textured (Cur_N and Cur_S) profiles.

Porewater profiles were collected at each of the four sampling locations on two separate occasions: the first 5 mo after rewetting ( January 2010) and another 24 mo following rewetting (August 2011). Data collected 5 mo following rewetting are collectively referred to as post-rewet/+5 and data collected 24 mo following rewetting are collectively referred to as post-rewet/+24.

Fig. 1. Top left: the study area within the murray–darling Basin. Top right: the location of the study area in relation to the mouth of the murray River. Bottom: Lake Alexandrina and positions of the four sampling locations in the Finniss River (Fin) and Currency Creek (Cur) catchments.

www.agronomy.org • www.crops.org • www.soils.org 991

Porewater Sampling and CharacterizationDepth profiles of porewater chemistry were obtained

from each sampling location using in situ equilibrium dialysis membrane samplers, commonly known as peepers (Hicks et al., 2009a; Johnston et al., 2009a; Teasdale et al., 1995; van Oploo et al., 2008) (Supplementary Fig. S1). The peepers were constructed of poly(methyl methacrylate) (Perspex) (Carignan et al., 1994). The sampling interval covered 35 cm split into 3 cm in the overlying water and the remaining 32 cm in the subaqueous soil with a potential sampling resolution of 1 cm (i.e., 1-cm spacing between sampling chambers). The peepers used were open on both sides and similar in design to Hesslein-style peepers (Hesslein, 1976; Webster et al., 1998), allowing reduced equilibrium times. Each sampling chamber contained approximately 3.2 mL of sample.

The peepers were prepared for assembly immediately before deployment by cleaning in detergent solution followed by a 10% v/v HNO3 bath (72 h) and a thorough rinse with deionized water (type 1 reagent grade) (American Public Health Assoc., 2012). To assemble the peepers, each chamber was filled with deoxygenated and deionized water and sealed with a permeable, inert polysulfone membrane (pore size 0.45 mm). Care was taken to ensure gas bubbles were excluded. The membrane was then fixed in place with a separate Perspex cover plate and nylon screws. Once assembled, peepers were submerged under deoxygenated and deionized water and purged with N2 for 48 h. Deoxygenation took place in a portable acid-washed, plasticizer-free polyvinylchloride vessel capable of maintaining a deoxygenated environment during transport to the field. Before sealing the transport vessel the flow rate of the N2 was increased to exclude any oxygen from the headspace.

Peepers were inserted into the subaqueous soils using a specialized apparatus that minimizes soil disturbance, ensures a vertical deployment, and accurately deploys the peeper with a predetermined number of chambers in the overlying water column. At Finniss River sampling locations, where transhorizon cracking was present, the peeper was inserted into the soil in the center of a pedon. Peepers were deployed in duplicate at each sampling location to increase sample volume. Samples from the corresponding chambers of each duplicate peeper were combined, effectively decreasing variability. Duplicate peepers were placed ~0.5 m apart to avoid equilibrium interferences (Harper et al., 1997). Peepers were deployed on two separate occasions: the first for 33 d from 21 December 2009 to 22 January 2010 and the second for 34 d from 22 July 2011 to 25 August 2011. An equilibration period of ≥30 d is more than sufficient for the double-sided peeper design used in this study and the soil physical characteristics of the sampling locations (Carignan et al., 1985; Grigg et al., 1999; Teasdale et al., 1999). To avoid any site disturbance that may have occurred during the first deployment, the second peeper deployment was separated by ~2 m from the site of the first.

Peepers were removed one at a time from the subaqueous soil and sample collection was completed within 5 to 10 min for each peeper to minimize the transformation of redox-sensitive species. Sample was extracted from the peeper with a 5-mL micropipette, transferred to a 10-mL screw top vial and chilled to 4°C until analysis. Anoxic cells were also sampled first to further minimize

transformation of reduced redox-sensitive species. The next day, a subsample was removed for immediate analysis of pH, electrical conductivity, and alkalinity or acidity. Measurements of pH and alkalinity or acidity were made using an Orion 960 autotitrator and Orion micro pH electrode (Thermo Scientific). Total acidity was also calculated from individual acidic cations: one mole of total dissolved Fe was assumed to contribute 2 moles of H+ acidity; total dissolved Al, 3 moles of H+ acidity; and total dissolved Mn, 2 moles of H+ acidity. The remaining sample was preserved with HNO3 and stored at 4°C for subsequent multielement analysis. Major-element concentrations were determined by inductively coupled plasma atomic emission spectrometry and trace elements by inductively coupled plasma mass spectrometry. The SO4

2− and Cl− concentrations were measured by ion chromatography. For quality assurance, two additional peepers were prepared and transported to the field to act as field blanks. The field blanks were prepared, sampled, and analyzed in an identical manner to the deployed peepers. Analysis of field blanks showed no contamination during the preparation, deployment, retrieval, and sample preparation processes.

ResultsGeneral PropertiesFinniss River Sampling Locations (Clayey)

Before the rewetting event (pre-rewet/-5), soil profiles taken from the Finniss catchment (Fin_N and Fin_S) were classified as Hydraquentic Sulfaquepts (Soil Survey Staff, 2014) (Supplemental Table S1). Currently, no subgroup exists in soil taxonomy to correctly classify the Finniss River soils following their rewetting (post-rewet/+5 and post-rewet/+24). They can be best described as subaqueous soils with sulfuric horizons. Fitzpatrick and Grealish (Fitzpatrick, unpublished data, 2015) have suggested the term Hydraquentic Sulfowassepts to describe these soils (see Supplemental Table S1). Soil texture at the Fin_N and Fin_S sampling locations was dominantly clayey, although Fin_S had a higher percentage clay content compared with Fin_N (Fig. 2). Both sampling locations in the Finniss River catchment exhibited deep but partially infilled transhorizon desiccation cracks of 20 and 10 cm deep at Fin_N and Fin_S, respectively. Desiccation cracks remained present following rewetting (post-rewet/+5 and post-rewet/+24). Before the drought, the potential IASS at Fin_N and Fin_S were very soft with a low load-bearing capacity (n value >1) (Soil Survey Staff (2014). Cracking at these sample locations was the result of soil moisture loss during the drought and is not related to shrinkage and swelling of clays. Once formed, this type of cracking is considered permanent (Smiles, 2000; Soil Survey Staff, 2014). Color photographs of the soil profiles before rewetting can be found in Fitzpatrick et al. (2011).

Both Finniss River sampling locations had sulfuric horizons with visibly present schwertmannite and jarosite or natrojarosite before rewetting. Schwertmannite was observed as a surface crust and on ped faces in the top 2 to 3 cm of the profiles. Jarosite or natrojarosite was typically observed deeper in the profile on ped faces and near old root channels. Jarosite or natrojarosite was found to persist following rewetting and throughout the study period. The visual observation of jarosite or natrojarosite

992 Journal of Environmental Quality

persisting after 24 mo of subaqueous conditions was supported by measured redox potential (Eh) and pH data that indicated a redox environment within the stability field for jarosite.

Surface water pH was circumneutral above the soil–water interface (SWI) for both post rewetting sampling events (Post-rewet/+5 and Post-rewet/+24) at both Finniss River sampling locations (Fig. 2). Following 5 mo of subaqueous conditions, soil porewater pH below the SWI decreased with depth to values approaching pH 3.5 or less. The down-profile pH change at Fin_S was abrupt at ~4 cm below the SWI, decreasing by ~2.5 pH units over less than 2 cm. The pH decreased more gradually (over ~15 cm) from circumneutral surface water to acidic (pH <3.5) soil porewater at Fin_N.

After a total of 24 mo of continuous subaqueous conditions, the majority of the profile at both Finniss River sampling locations remained acidic (pH <4) with only small increases in pH (~0.5 units) being observed. A larger increase in pH (~2–2.5 pH units) was observed at Fin_S but was confined to the uppermost 10-cm portion of the profile.

Alkalinity in the surface water at sampling location Fin_N was typical of historic levels (~2.5 mmol HCO3

− L−1) (Post-rewet/+5) and was present in the porewaters of Fin_N to a depth of 13 cm below the SWI. In contrast, little alkalinity existed in the surface water at Fin_S (≤1 mmol HCO3

− L−1) and porewater alkalinity was exhausted before a depth of 3 cm below the SWI. Acidity values at both sampling locations increased quickly once alkalinity was depleted. Acidity levels were very high, reaching maximum values of 84 and 43 mmol H+ L−1 near the bottom of the profile at Fin_N and Fin_S, respectively.

Alkalinity in the surface water of both Finniss River sampling locations was substantially lower following a further 19 mo of subaqueous conditions. Decreases in porewater acidity 24 mo after rewetting was negligible at Fin_S and only moderate at Fin_N. Overall acidity levels remained high, and soil porewater remained highly buffered at low pH values.

Five months after rewetting, porewater acidity at Finniss River sampling locations was only in part due to low pH (H+ ion). Acidity was dominated by total dissolved Fe (Fe2+ > Al3+ >> Mn2+ » H+). Acidity became further dominated by total dissolved Fe after a further 19 mo of subaqueous conditions (Fe2+ >> Al3+ » Mn2+ » H+) due to large decreases in the concentrations of Al, Mn, and H+ (increased pH) combined with smaller decreases in total dissolved Fe concentrations.

After 5 mo of subaqueous conditions, porewater Cl− concentrations in the upper 10 cm of the profile decreased to

that of the surface water at both Finniss River sampling locations (Fig. 2). Surface water was slightly fresher at Finniss River sampling locations (~50 mmol L−1) than those at Currency Creek (~100 mmol L−1). Below a depth of 10 cm there was a linear down-profile increase in porewater Cl− to maxima of ~75 and ~140 mmol L−1 for Fin_N and Fin_S, respectively.

Fig. 2. down-profile morphological characteristics and trends in soil porewater pH, Cl−, and alkalinity or acidity of from Finniss River (Fin) and Currency Creek (Cur). Alkalinity (HCO3

−) and acidity (H+) are plotted to the left and right of the vertical line, respectively. The soil–water interface is at 0-cm depth. Post-rewet/+5 (open circles). Post-rewet/+24 (solid triangles). morphological characteristics before rewetting from left to right: texture, redoximorphic features (Jar. = Jarosite, Sch. = Schwertmannite), soil cracking, and, matrix color (moist munsell color).

www.agronomy.org • www.crops.org • www.soils.org 993

Surface water Cl− concentrations reduced to ~12 mmol L−1 at all of the sampling locations following a further 19 mo of continued subaqueous conditions. Porewater Cl− concentrations in the upper 10 to 15 cm of the profiles also decreased. Below 15 cm, minor decreases in Cl− were observed at Fin_N and concentrations remained steady at Fin_S

Currency Creek Sampling Locations (Sandy)Soil profiles in the Currency Creek catchment (Cur_N and

Cur_S) comprised loose, loamy sand overlying sandy clay. The depth of loamy sand was approximately 5 and 35 cm for Cur_N and Cur_S, respectively (Fig. 2). Currency Creek profiles (Cur_N and Cur_S) were classified as Typic Sulfaquepts (Soil Survey Staff, 2014) before the rewetting event soil profiles (pre-rewet/-5) (Supplemental Table S1). Following rewetting (post-rewet/+5 and post-rewet/+24) the sulfuric horizons persisted, and they were best described as subaqueous soils with sulfuric horizons. No subgroup currently exists in soil taxonomy to correctly classify the rewetted Currency Creek soils, but Fitzpatrick and Grealish (Fitzpatrick, unpublished data, 2015) have suggested the term Typic Sulfowassepts to describe these soils (Supplemental Table S1).

Sulfuric horizons with visibly present jarosite existed at both Currency Creek sampling locations before rewetting (Fig. 2). Schwertmannite was also observed together with the jarosite or natrojarosite at Cur_N. Schwertmannite and jarosite or natrojarosite was typically observed on the surfaces of old root channels. As with the Finniss River sampling locations, jarosite or natrojarosite remained visibly present after 24 mo of subaqueous conditions. Schwertmannite, observed at Cur_N before rewetting, remained visually observable following 5 mo of subaqueous conditions but was no longer visually observable after 24 mo of subaqueous conditions.

Surface water pH was circumneutral above the SWI at Cur_N and Cur_S (Fig. 2). Below the SWI, soil porewater pH decreased abruptly to values approaching pH 4 or less. The abrupt down-profile pH change after 5 mo of subaqueous conditions, whereby porewater became severely acidic, occurred ≤1 cm below the SWI interface at Cur_N and ~5 cm below the SWI at Cur_S.

After a total of 24 mo of continuous subaqueous conditions, both Currency Creek sampling locations remained mostly acidic (pH <4), with only small increases in pH (~0.5 units) being observed. Larger increases in pH (~2–2.5 pH units) were observed at Cur_N but were confined to the top 7 cm of the soil profile and the bottom 4 cm of the sampled range (29–32 cm below the SWI). A marked decrease of 2.5 pH units was also observed at the bottom of the sampled profile at Cur_S after 24 mo of subaqueous conditions, however, this likely represents spatial heterogeneity, whereby the first peeper installation encountered sulfidic material in the underlying layer and the second installation did not.

Following 5 mo of subaqueous conditions, alkalinity in the surface water at Cur_S was typical of historic levels (~2.5 mmol HCO3

− L−1) (Post-rewet/+5) (Fig. 2). In contrast, alkalinity concentrations were very low (≤1 mmol HCO3

− L−1) in the surface water at Cur_N. Porewater alkalinity was also present in the porewater of both Currency Creek sampling locations but was exhausted at depth of 3 cm below the SWI. Acidity values increased quickly once alkalinity was depleted, reaching

high concentrations in midprofile at Cur_N (maximum 55 mmol H+ L−1) and Cur_S (maximum 72 mmol H+ L−1).

Following 24 mo of subaqueous conditions alkalinity in the surface waters above the SWI of both Currency Creek sampling locations were ≤1 mmol HCO3

− L−1 and no alkalinity was present in the porewater below the SWI. Moderate decreases in acidity (up to 30 mmol H+ L−1) were observed in the porewater of the Currency Creek sampling locations, however, overall acidity levels remained high, with maximum values of 31 and 60 mmol H+ L−1 for Cur_N and Cur_S, respectively.

Five months after rewetting porewater acidity was mainly comprised of total dissolved Fe at Cur_N (Fe2+ > Al3+ >> Mn2+ » H+) (Fig. 2). After a further 19 mo of subaqueous conditions, acidity became further dominated by total dissolved Fe at Cur_N (Fe2+ >> Al3+ » Mn2+ » H+) due to decreases in the concentrations of Al, Mn, and H+ (increased pH). At Cur_S, high concentrations of Al meant that acidity was dominated by soluble Al3+ species 5 mo after rewetting (Al3+ > Fe2+ >> Mn2+ » H+). Aluminum and Fe contributed approximately equally to the acidity at Cur_S after 24 mo of subaqueous conditions (Al3+ » Fe2+ >> Mn2+ » H+).

Surface water Cl− concentrations were higher at Currency Creek sampling locations (~100 mmol L−1) than at the Finniss River sampling locations (~50 mmol L−1) following 5 mo of subaqueous conditions (Fig. 2). Down-profile Cl− concentrations increased at Cur_N (from 100 to 120 mmol L−1) and decreased at Cur_S (from 100 to 70 mmol L−1). However, down-profile variances between surface water and porewater Cl− concentrations were much smaller at Currency Creek sampling locations than those at the Finniss River sampling locations 5 mo after rewetting.

Following a further 19 mo of continuous subaqueous conditions, surface water Cl− concentrations reduced to ~12 mmol L−1 at both Currency Creek sampling locations. Porewater Cl− concentrations also decreased in the upper 10 to 15 cm of the both Currency Creek sampling locations. Less significant decreases in Cl− concentrations were observed at Cur_N below 15 cm. After 24 mo of subaqueous condition, Cl− concentrations at Cur_S increased by ~20 mmol L−1 15 cm below the SWI and ~50 mmol L−1 30 cm below the SWI.

Iron and SulfateFinniss River Sampling Locations (Clayey)

Total dissolved Fe in the surface waters of Finniss River sampling locations was generally very low (≤4 mmol L−1) or below the detection limit (0.009 mmol L−1) after 5 mo, and less than detection after 24 mo of subaqueous conditions (Fig. 3).

Total dissolved Fe in the porewater increased with depth to a maximum near the bottom of the sampled depth range (post-rewet/+5). At Fin_N, Fe remained low (1–5 mmol L−1) to a depth of 5 cm before increasing to a maximum of ~45 mmol L−1 at 14 cm; concentrations then remained approximately constant over the rest of the sampled depth range. Maximum Fe concentrations were much higher at Fin_N (~45 mmol L−1) than at any other of the three sampling locations, which had maximum concentrations in the range 10 to 20 mmol L−1. At Fin_S, Fe concentrations increased with depth in a more linear fashion except for a large spike in concentration observed

994 Journal of Environmental Quality

2  cm below the SWI. The down profile trends of SO4

2− concentrations followed those of Fe, although SO4

2− concentrations were higher (Fig. 3). Sulfate was also elevated relative to Cl− and coincided with low pH values. Porewater Fe and SO4

2− decreased during the time between the first and second sampling occasions. Larger decreases were observed in the upper part of the profiles and generally coincided with decreases in Cl− and increases in pH.

Currency Creek Sampling Locations (sandy)Total dissolved Fe concentrations after 5

mo of subaqueous conditions were generally very low (≤1 mmol L−1) or below the detection limit (0.009 mmol L−1) in the surface waters of both Currency Creek sampling locations (Fig. 3). Surface water located just above the SWI (1 cm) had slightly higher Fe concentrations (~5 mmol L−1) at Cur_N. After 24 mo of subaqueous conditions, Fe concentrations in the surface waters were below detection limit at all sampling locations.

The down-profile changes to porewater Fe concentrations 5 mo after rewetting were similar for both Currency Creek sampling locations. Unlike the Finniss River sampling locations, where maximum Fe concentrations occurred near the bottom of the sampling depth, Fe concentrations in the Currency Creek profiles increased quickly once soil pH became acidic (pH <4), and maximum Fe concentrations (~20 and 10 mmol L−1 for Cur_N and Cur_S, respectively) were reached at a much shallower depth (~10 cm). Below this depth, Fe concentrations decreased, reaching concentrations close to those near the SWI at depth of 30 cm. Porewater Fe and SO4

2− concentrations decreased at both Currency Creek sampling locations during the time between the first and second sampling occasions with the largest decreases observed in the upper 15 cm of the profiles (Fig. 3).

Trace Element BehaviorFinniss River Sampling Locations (Clayey)

On both sampling occasions, the surface water and porewater in the upper part of the Finniss River profiles (~5–15 cm below the SWI) contained much lower concentrations of trace metals than in the bottom half of profile (Fig. 4). Trace-metal concentrations below ~10 cm increased with depth reaching maximum concentrations between one and three orders of magnitude greater by the bottom of the sampled depth range. Following a further 19 mo of subaqueous conditions considerable decreases in the concentrations Al, Cr, Ni, and Zn were observed, with smaller decreases in Mn concentrations.

Currency Creek Sampling Locations (sandy)The surface waters of both Currency Creek sampling locations

had low concentrations of trace metals on both sampling occasions (Post-rewet/+5 and Post-rewet/+24) (Fig. 4). Similar to the Finniss River profiles, trace-metal concentrations increased

Fig. 3. down-profile soil porewater trends of total dissolved Fe and SO42− concentrations

and their Cl− ratios of Finniss River (Fin) and Currency Creek (Cur). Note differing x-axis scales for some plots. Post-rewet/+5 (open circles). Post-rewet/+24 (solid triangles).

www.agronomy.org • www.crops.org • www.soils.org 995

below the SWI as porewater became acidic (pH <4), reaching maximum concentrations one to three orders of magnitude greater than those found in the surface water. In contrast to the Finniss River sampling locations, maximum trace-metal concentrations occurred in the middle of the sampled profiles. Trace-metal concentrations below the maximum decreased with increasing depth, generally at a rate slower than the rate of increase above the concentration peak. By the bottom of the sampled range, trace-metal concentration approached (Cur_N) or reached (Cur_S) concentrations similar to those found in the surface water. Following a further 19 mo of subaqueous conditions, considerable decreases in the concentrations of all trace metals were observed, with the largest decreases occurring midprofile.

DiscussionSoil Structure Impacts on Infiltration and Rate of Recovery

The removal of soil acidity followed by the improvement of soil pH toward circumneutral conditions is an important first requirement toward the complete recovery (i.e., transformation from a sulfuric horizon to sulfidic material). During the initial phase of recovery and in the absence of in situ microbial alkalinity generation such as the reduction of Fe and SO4

2− (see later discussion), soil acidity is removed by dilution or neutralization via surface water infiltration and downward advective flow. Hence, the presence or absence of infiltration and mixing zones is a key control in determining the likely rate of initial recovery. Due to its conservative chemical behavior, Cl− was used as a tracer to determine presence or absence of infiltration and develop a conceptual model for the initial phase of recovery in dominant clayey-textured and dominantly sandy-textured profiles (Fig. 5). Surface water Cl− concentrations were lower on the second sampling occasion due to the continued high postdrought freshwater flows and local winter rainfall between the first and second sampling occasions (Fig. 2). Therefore, any observed decrease in porewater Cl− concentrations between the first and second sampling occasions is indicative of surface water inputs.

At Fin_N, porewater Cl− decreased between the first and second sampling occasions indicating a recharge mixing zone at least as deep as the sampled depth range (0–32 cm below SWI). Decreases in Cl− between the first and second sampling occasions were similar over the sampled depth range, indicating that infiltration may be occurring through ped faces as well as through the SWI. If infiltration were occurring dominantly through the SWI, decreases in Cl− concentrations would be largest near the SWI and get progressively small with increasing depth. Deep desiccation cracking (20 cm below SWI) was recorded at Fin_N, however, the oxidized sulfuric horizon at Fin_N reached a depth of 50 cm. Macropores below the depth of cracking recorded and extending beyond the 32-cm peeper sampling depth can explain the relative rapid (<24 mo) infiltration of surface water into a clay-textured profile over the entire sampled depth range.

At Fin_S, evidence of surface water infiltration was only observed in the top 15 cm of the profile, coinciding approximately with the depth of cracking. Porewater Cl− in the top 8 cm of the profile was in equilibrium with the surface water 24 mo after rewetting. Evidence of advective transport

below 15 cm (below depth of cracking) was absent after 24 mo of subaqueous conditions and is consistent with the shallower cracking at Fin_S and the low hydraulic conductivity that would be expected for massive structured clays. These results indicate that the depth of cracking is a major control on the initial rate of recovery in dominantly clay-textured soils with low hydraulic conductivity (i.e., Fin_N and Fin_S). The different behavior of the Finniss River sampling locations—two spatially close profiles with similar starting conditions and physical properties—also illustrates the heterogeneity of IASS and the difficulties of extrapolating expected recovery outcomes to scales greater than site specific.

At Cur_N surface water movement occurred to a depth greater than or equal to the sampled depth range (32 cm below SWI) between the first and second sampling occasions. Changes in porewater Cl− concentrations at Cur_N are indicative of infiltration through the SWI, with the largest decreases in Cl− concentrations after 24 mo of rewetting occurring closest to the SWI. The advective flow behavior was similar at Cur_S and Cur_N between the first and second sampling occasions, but below 10 cm, Cl− concentrations increased between the first and second sampling occasions. This increase is expected to be due to local spatial heterogeneity and likely represents the lateral or upward intrusion of more saline groundwater. The observation is also consistent with the shallow (29 cm below SWI) circumneutral layer containing sulfidic material encountered during the first sampling occasion at the Cur_S sampling location (Fig. 2).

Circumneutral pH surface waters were able to infiltrate the acidic profiles of all sampling locations, albeit to different degrees. Mixing between fresh surface water and a deeper, more saline groundwater explains the observed Cl− profiles, with the largest decreases in soil acidity occurring where mixing was greatest. The loss of soil acidity and increase in soil pH that occurred following the first 24 mo of subaqueous conditions is an indicator of the estimated rate of initial recovery. The slow rate of mixing between circumneutral surface waters and acidic porewaters, combined with high acidity concentrations, persistence of retained acidity in the form of acidic Fe(III)-SO4

2− minerals, and low surface water alkalinity concentrations typically associated with postdrought inflowing flood waters all indicate that initial recovery is likely to take several years.

Changes in Iron and Sulfate ConcentrationsAt all sampling locations, porewater SO4

2− concentrations were elevated relative to Cl−, indicating that soil acidification was a result of pyrite oxidation during desiccation. At Fin_S and Cur_N, the decreases in total dissolved Fe and SO4

2− concentrations relative to Cl− between the first and second sampling occasions were generally what would be expected if these behaved conservatively (Fig. 6). The large spike in total dissolved Fe concentrations observed 2 cm below the SWI at Fin_S (post-rewet/+5) likely represents the reductive dissolution of Fe(III) minerals that had accumulated near the surface. Certainly, the profile exhibited red-brown mottling typical of Fe(III) oxides at this depth. The upward transport of Fe(II)-rich porewaters, and their subsequent oxidation when they reach a more oxic environment nearer the SWI, has been suggested to lead to the accumulation of Fe(III) minerals near the SWI (Anderson and

996 Journal of Environmental Quality

Schiff, 1987; van Breemen, 1975). It is probable that the same process is occurring in this profile.

At sampling locations Fin_N and Cur_S, Fe and SO42−

concentrations were depleted beyond what would be expected conservatively (Fig. 6), suggesting the presence of a

Fig. 4. down-profile soil porewater trends of Al, Cr, mn, Ni, and Zn of Finniss River (Fin) and Currency Creek (Cur). dashed line represents guideline trigger value. Post-rewet/+5 (open circles). Post-rewet/+24 (solid triangles).

www.agronomy.org • www.crops.org • www.soils.org 997

nonconservative behavior that removes Fe and SO42− from

solution. In rewetted soil profiles containing sulfidic material, the consumption of Fe and SO4

2− would usually be expected to be a result of their incorporation into iron sulfide minerals (pyrite, mackinawite) (Postma and Jakobsen, 1996). However, the microbial reduction of SO4

2− requires relatively strong reducing conditions in contrast to the high Eh–low pH redox environment present in these profiles (Fig. 4). Although considered an unexpected occurrence for subaqueous soil with a sulfuric horizon, it is possible that in these persisting low pH–high Eh redox environments the loss of Fe and SO4

2− from solution may be in part due to the formation of new Fe(III)-SO4

2− minerals (e.g., schwertmannite). As well as low pH, limitation of electron donors (e.g., organic carbon), elevated concentrations of metals toxic to bacteria, and competition from other bacteria communities such as Fe

reducers can also inhibit microbial SO42− reduction. (Kelly

et al., 1995; Koschorreck, 2008). We suspect that low pH is the dominant factor limiting microbial SO4

2− reduction and subsequent monosulfide and pyrite formation during the initial recovery of these soils. We do not suspect that the microbial reduction of SO4

2− during the assessed period was carbon limited. However, once soil acidity has decreased, SO4

2− reduction may become carbon limited (Berner, 1984; Sullivan et al., 2010).

Changes in Contaminant ConcentrationsAt all sampling locations, the concentrations of assessed

trace metals decreased between the first and second sampling occasions (Fig. 4). Relative to Cl, the decrease in trace-metal concentrations was much greater than would be expected for conservative behavior (Fig. 6). The solubility of Al, Mn,

Fig. 5. Conceptual process model illustrating initial recovery in the first 24 mo after rewetting for dominantly clay-textured and dominantly sand-textured profiles.

Fig. 6. Plots demonstrating nonconservative behavior (relative to Cl) of Fe, Al, Cr, and Ni of Finniss River (Fin) and Currency Creek (Cur). dashed line represents conservative behavior (i.e., Na). Left of dashed line represents a percentage loss beyond what would be expected conservatively. Right of line represents a percentage gain beyond what would be expected conservatively.

998 Journal of Environmental Quality

Ni, Cr, and Zn is highly dependent on solution pH (and for redox sensitive species also Eh), with solubility decreasing with increasing pH (Appelo and Postma, 2005). A strong negative correlation between pH and trace-metal concentrations (Fig. 7) supports that substantial decreases in the concentrations of some trace metals between the first and second sampling occasions is likely owing to pH-dependent immobilization. The almost vertical trend in trace-metal concentrations at ~pH 3 (Fig. 7) indicates a highly pH buffered (i.e., by jarosite or natrojarosite) system. Figure 2 also illustrates how small increases in porewater pH after 24 mo of continuous subaqueous conditions were sufficient to decrease the solubility of Al, Mn, Ni, Cr, and Zn and remove many of the cations from solution.

ConclusionsThe results of the study demonstrate that the introduction

of subaqueous conditions did not result in rapid recovery of ASS with sulfuric horizons. The soils remained largely acidic over a 24-mo time period, with recovery to previous anoxic, circumneutral conditions confined to the uppermost parts of the profiles. The approximate timescales involved for the complete recovery (i.e., neutralization or removal of acidity and the

reestablishment of reducing conditions) of IASS with sulfuric horizons rewetted by a freshwater body are, therefore, very long.

Low pH conditions and high porewater metal concentrations persisted within the soil profiles for at least 5 mo after rewetting. Following a further 19 mo of subaqueous conditions there was an observed improvement in water quality corresponding to substantial decreases in trace-metal concentrations associated with small increases in pH. The highest increase in pH was generally confined to the uppermost 5 cm of the profiles and only small 0.5 to 1 unit increases were observed over the rest of the profile. Soil porewater remained heavily buffered at these low pH values.

The persistent low pH–high Eh conditions in the sampled soil profiles are likely slowing the microbial reduction of Fe and SO4

2− and preventing the concurrent production of additional in situ alkalinity via proton consumption in the formation of pyrite. For anoxic, circumneutral pH conditions to migrate downward from the uppermost part of the profiles, existing porewater acidity plus any acidity retained in the persisting acidic Fe(III)-SO4

2− minerals not yet released must be neutralized. Currently, the very low amounts of alkalinity in the surface water are likely the only method of neutralization easily available and its supply is limited by infiltration through

Fig. 7. Relationship between trace-metal concentration and porewater pH for 5 mo after rewetting and 24 mo after rewetting.

www.agronomy.org • www.crops.org • www.soils.org 999

the SWI or, in the case of Finniss River sampling locations, also the ped walls. Therefore, recovery is likely to continue at a slow rate until the predominately top-down migration of alkalinity from the surface water sufficiently raises the pH to levels where the microbial formation of pyrite from Fe(III) and SO4

2− begins to actively produce alkalinity in situ.

The possibility of a slow, multiyear recovery for IASS with sulfuric horizons rapidly overtopped by a freshwater body is an important finding of this study as the persistence of low-pH conditions is likely to present a higher ecotoxicological risk than if recovery was rapid. Continued examination of these profiles as they recover would further add to the understanding of the behaviors of IASS with sulfuric horizons rewetted in inland freshwater systems and, therefore, would be worthy of further investigation.

Supplemental MaterialSupplemental material associated with this article can be

found online.

AcknowledgmentsThis work was largely funded by the Department of Environment and Natural Resources, Murray Futures Lower Lakes and Coorong Recovery Acid Sulfate Soils Research Program, South Australia. We would like to acknowledge the support and comments of Richard Merry, Sonia Grocke, Gerard Grealish, and Nilmini Jayalath. We also thank the anonymous reviewers for their constructive and valuable additions to this paper.

ReferencesAmerican Public Health Association. 2012. Standard Methods for the

Examination of Water and Wastewater. 22nd ed. American Public Health Association, American Water Works Association, and Water Environment Federation, Washington, DC.

Anderson, R.F., and S.L. Schiff. 1987. Alkalinity generation and the fate of sulfur in lake sediments. Can. J. Fish. Aquat. Sci. 44:s188–s193. doi:10.1139/f87-294

Appelo, C.A.J., and D. Postma. 2005. Geochemistry, groundwater and pollution. 2nd ed. Balkema Publications, Leiden, the Netherlands.

Astrom, M., and J. Astrom. 1997. Geochemistry of stream water in a catchment in Finland affected by sulphidic fine sediments. Appl. Geochem. 12:593–605. doi:10.1016/S0883-2927(97)00016-4

Berner, R.A. 1984. Sedimentary pyrite formation: An update. Geochim. Cosmochim. Acta 48:605–615. doi:10.1016/0016-7037(84)90089-9

Burton, E.D., R.T. Bush, L.A. Sullivan, S.G. Johnston, and R.K. Hocking. 2008. Mobility of arsenic and selected metals during re-flooding of iron- and organic-rich acid-sulfate soil. Chem. Geol. 253:64–73. doi:10.1016/j.chemgeo.2008.04.006

Burton, E.D., I.R. Phillips, and D.W. Hawker. 2005. Geochemical partitioning of copper, lead, and zinc in benthic, estuarine sediment profiles. J. Environ. Qual. 34:263–273. doi:10.2134/jeq2005.0263

Carignan, R., S. St. Pierre, and R. Gachter. 1994. Use of diffusion samplers in oligotrophic lake sediments: Effects of free oxygen in sampler material. Limnol. Oceanogr. 39:468–474. doi:10.4319/lo.1994.39.2.0468

Carignan, R., F. Rapin, and A. Tessier. 1985. Sediment porewater sampling for metal analysis: A comparison of techniques. Geochim. Cosmochim. Acta 49:2493–2497. doi:10.1016/0016-7037(85)90248-0

Creeper, N., R.W. Fitzpatrick, and P. Shand 2013. The occurrence of inland acid sulphate soils in the floodplain wetlands of the Murray-Darling Basin, Australia, identified using a simplified incubation method. Soil Use Manage.29:130–139. doi:10.1111/sum.12019

Dent, D.L. 1986. Acid sulphate soils: A baseline for research and development. ILRI Pub. No. 39. International Institute for Land Reclamation and Improvement, Wageningen, the Netherlands.

Earth Systems. 2010. Quantification of acidity flux rates to the Lower Murray Lakes. Earth Systems Consulting Pty. Ltd. for the South Australia Dep. of Environ. and Nat. Resour. Adelaide.

Fitzpatrick, R.W., A.K.M. Baker, P. Shand, R.H. Merry, G. Grealish, and L.M. Mosley. 2012. A modern soil–landscape characterization approach to reconstructing and predicting pedogenic pathways of inland acid sulfate soils. Geol. Surv. Finl. Bull. 56:47–50.

Fitzpatrick, R.W., G.J. Grealish, P. Shand, and N.L. Creeper. 2011. Monitoring and assessment of reflooded acid sulfate soil materials in Currency Creek and Finniss River Region, South Australia. Client Report R-325-8-6. CSIRO Sustainable Agriculture National Research Flagship. Adelaide.

Fitzpatrick, R.W., G. Grealish, P. Shand, S.L. Simpson, R.H. Merry, and M.D. Raven. 2009a. Acid sulfate soil assessment in Finniss River, Currency Creek, Black Swamp, and Goolwa Channel, South Australia. Science Report 26/09. CSIRO Land and Water, Adelaide.

Fitzpatrick, R.W., G. Grealish, P. Shand, B.P. Thomas, R.H. Merry, N.L. Creeper, et al. 2009b. Preliminary risk assessment of acid sulfate soil materials in the Currency Creek, Finniss River, Tookayerta Creek and Black Swamp region, South Australia. Science Report 01/09. CSIRO Land and Water, Adelaide.

Fitzpatrick, R.W., P. Shand, and R.H. Merry. 2009c. Acid sulfate soils. In: J.T. Jennings, editor, Natural history of the Riverland and Murraylands. Royal Society of South Australia, Adelaide, South Australia. p. 65–111.

Glover, F., K.L. Whitworth, P. Kappen, D.S. Baldwin, G.N. Rees, J.A. Webb, et al. 2011. Acidification and buffering mechanisms in acid sulfate soil wetlands of the Murray–Darling Basin, Australia. Environ. Sci. Technol. 45:2591–2597. doi:10.1021/es103535k

Grigg, N.J., I.T. Webster, and P.W. Ford. 1999. Pore-water convection induced by peeper emplacement in saline sediment. Limnol. Oceanogr. 44:425–430. doi:10.4319/lo.1999.44.2.0425

Hall, K.C., D.S. Baldwin, G.N. Rees, and A.J. Richardson. 2006. Distribution of inland wetlands with sulfidic sediments in the Murray–Darling Basin, Australia. Sci. Total Environ. 370:235–244. doi:10.1016/j.scitotenv.2006.07.019

Harper, M.P., W. Davison, and W. Tych. 1997. Temporal, spatial, and resolution constraints for in situ sampling devices using diffusional equilibration: Dialysis and DET. Environ. Sci. Technol. 31:3110–3119. doi:10.1021/es9700515

Hesslein, R.H. 1976. In situ sampler for close interval pore water studies. Limnol. Oceanogr. 21:912–914. doi:10.4319/lo.1976.21.6.0912

Hicks, W.S., G.M. Bowman, and R.W. Fitzpatrick. 2009a. Effect of season and landscape position on the aluminium geochemistry of tropical acid sulfate soil leachate. Aust. J. Soil Res. 47:137–153. doi:10.1071/SR06106

Hicks, W.S., N.L. Creeper, J. Hutson, R.W. Fitzpatrick, S. Grocke, and P. Shand. 2009b. The potential for contaminant mobilisation following acid sulfate soil rewetting: Field experiment. Commonwealth Scientific and Industrial Research Organisation Land and Water for the South Australia Dep. of Environ. and Nat. Resour., Adelaide. CSIRO, Adelaide.

Hicks, W.S., R.W. Fitzpatrick, S. Lamontagne, and S. Rogers. 2003. Risks of water quality degradation during the remediation of floodplain salinity in the River Murray. In: I.C. Roach, editor, Advances in regolith. CRC LEME, Adelaide, Australia. p. 178–181.

Johnston, S.G., E.D. Burton, R.T. Bush, A.F. Keene, L.A. Sullivan, D. Smith, et al. 2010a. Abundance and fractionation of Al, Fe and trace metals following tidal inundation of a tropical acid sulfate soil. Appl. Geochem. 25:323–335. doi:10.1016/j.apgeochem.2009.11.015

Johnston, S.G., E.D. Burton, A.F. Keene, R.T. Bush, L.A. Sullivan, and L. Isaacson. 2009a. Pore water sampling in acid sulfate soils: A new peeper method. J. Environ. Qual. 38:2474–2477. doi:10.2134/jeq2009.0135

Johnston, S.G., R.T. Bush, L.A. Sullivan, E.D. Burton, D. Smith, M.A. Martens, et al. 2009b. Changes in water quality following tidal inundation of coastal lowland acid sulfate soil landscapes. Estuar. Coast. Shelf Sci. 81:257–266. doi:10.1016/j.ecss.2008.11.002

Johnston, S.G., A.F. Keene, E.D. Burton, R.T. Bush, and L.A. Sullivan. 2012. Quantifying alkalinity generating processes in a tidally remediating acidic wetland. Chem. Geol. 304–305:106–116. doi:10.1016/j.chemgeo.2012.02.008

Johnston, S.G., A.F. Keene, E.D. Burton, R.T. Bush, L.A. Sullivan, A.E. McElnea, et al. 2010b. Arsenic mobilization in a seawater inundated acid sulfate soil. Environ. Sci. Technol. 44:1968–1973. doi:10.1021/es903114z

Johnston, S.G., A.F. Keene, R.T. Bush, E.D. Burton, L.A. Sullivan, D. Smith, et al. 2009c. Contemporary pedogenesis of severely degraded tropical acid sulfate soils after introduction of regular tidal inundation. Geoderma 149:335–346. doi:10.1016/j.geoderma.2008.12.013

Johnston, S.G., A.F. Keene, R.T. Bush, L.A. Sullivan, and V.N.L. Wong. 2011. Tidally driven water column hydro-geochemistry in a remediating acidic wetland. J. Hydrol. 409:128–139. doi:10.1016/j.jhydrol.2011.08.010

Keene, A.F., S.G. Johnston, R.T. Bush, E.D. Burton, and L.A. Sullivan. 2010. Reactive trace element enrichment in a highly modified, tidally inundated acid sulfate soil wetland: East Trinity, Australia. Mar. Pollut. Bull. 60:620–626. doi:10.1016/j.marpolbul.2010.02.006

1000 Journal of Environmental Quality

Keene, A.F., S.G. Johnston, R.T. Bush, L.A. Sullivan, E.D. Burton, A.E. McElnea, et al. 2011. Effects of hyper-enriched reactive Fe on sulfidisation in a tidally inundated acid sulfate soil wetland. Biogeochemistry 103:263–279. doi:10.1007/s10533-010-9461-2

Kelly, C.A., J.A. Amaral, M.A. Turner, J.W.M. Rudd, D.W. Schindler, and M.P. Stainton. 1995. Disruption of sulfur cycling and acid neutralization in lakes at low pH. Biogeochemistry 28:115–130. doi:10.1007/BF02180680

Koschorreck, M. 2008. Microbial sulphate reduction at a low pH. FEMS Microbiol. Ecol. 64:329–342. doi:10.1111/j.1574-6941.2008.00482.x

Lamontagne, S., W.S. Hicks, R.W. Fitzpatrick, and S. Rogers. 2006. Sulfidic materials in dryland river wetlands. Mar. Freshwater Res. 57:775–788. doi:10.1071/MF06057

Macdonald, B.C.T., J. Smith, A.F. Keene, M. Tunks, A. Kinsela, and I. White. 2004. Impacts of runoff from sulfuric soils on sediment chemistry in an estuarine lake. Sci. Total Environ. 329:115–130. doi:10.1016/j.scitotenv.2004.02.016

Mosley, L.M., P. Shand, and P. Self. 2014a. The geochemistry during management of lake acidification caused by the rewetting of sulfuric (pH <4) acid sulfate soils. Appl. Geochem. 41:49–61. doi:10.1016/j.apgeochem.2013.11.010.

Mosley, L.M., B. Zammit, A. Jolley, L. Barnett, and R.W. Fitzpatrick. 2014b. Monitoring and assessment of surface water acidification following rewetting of oxidised acid sulfate soils. Environ. Monit. Assess. 186:1–18. doi:10.1007/s10661-013-3350-9

Nyberg, M.E., P. Osterholm, and M.I. Nystrand. 2012. Impact of acid sulfate soils on the geochemistry of rivers in south-western Finland. Environ. Earth Sci. 66:157–168. doi:10.1007/s12665-011-1216-4

Nystrand, M.I., and P. Osterholm. 2013. Metal species in a Boreal river system affected by acid sulfate soils. Appl. Geochem. 31:133–141. doi:10.1016/j.apgeochem.2012.12.015

Portnoy, J.W., and A.E. Giblin. 1997a. Biogeochemical effects of seawater restoration to diked salt marshes. Ecol. Appl. 7:1054–1063. doi:10.1890/1051-0761(1997)007[1054:BEOSRT]2.0.CO;2

Portnoy, J.W., and A.E. Giblin. 1997b. Effects of historic tidal restrictions on salt marsh sediment chemistry. Biogeochemistry 36:275–303. doi:10.1023/A:1005715520988

Postma, D., and R. Jakobsen. 1996. Redox zonation: Equilibrium constraints on the Fe(III)/SO4-reduction interface. Geochim. Cosmochim. Acta 60:3169–3175. doi:10.1016/0016-7037(96)00156-1

Shand, P., R.H. Merry, S. Grocke, M. Thomas, R.W. Fitzpatrick, B. Thomas, et al. 2010. Water and soil quality impacts during reflooding of Nelwart Lagoon. CSIRO: Water for a Healthy Country National Research Flagship. ISSN: 1835-095X. CSIRO, Adelaide, Australia..

Simpson, S.L., R.W. Fitzpatrick, P. Shand, B.M. Angel, D.A. Spadaro, and L. Mosley. 2010. Climate-driven mobilisation of acid and metals from acid sulfate soils. Mar. Freshwater Res. 61:129–138. doi:10.1071/MF09066

Smiles, D.E. 2000. Hydrology of swelling soils: A review. Aust. J. Soil Res. 38:501–521. doi:10.1071/SR99098

Soil Survey Staff. 2014. Keys to soil taxonomy. 12th ed. USDA, NRCS, Washington, DC.

Sullivan, L.A., R.T. Bush, N.J. Ward, D.M. Fyfe, M. Johnston, E.D. Burton, et al. 2010. Lower Lakes laboratory study of contaminant mobilisation under seawater and freshwater inundation (long-term study). South Australian Environmental Protection Authority. Southern Cross GeoScience Tech. Rep. 1109. Southern Cross University, Lismore, NSW.

Teasdale, P.R., G.E. Batley, S.C. Apte, and I.T. Webster. 1995. Pore-water sampling with sediment peepers. Trac Trends Anal. Chem. 14:250–256.

Teasdale, P.R., S. Hayward, and W. Davison. 1999. In situ, high-resolution measurement of dissolved sulfide using diffusive gradients in thin films with computer-imaging densitometry. Anal. Chem. 71:2186–2191. doi:10.1021/ac981329u

van Breemen, N. 1975. Acidification and deacidification of coastal plain soils as a result of periodic flooding. Soil Sci. Soc. Am. J. 39:1153–1157. doi:10.2136/sssaj1975.03615995003900060035x

van Oploo, P., I. White, P. Ford, M.D. Melville, and B.C.T. Macdonald. 2008. Pore water chemistry of acid sulfate soils: Chemical flux and oxidation rates. Geoderma 146:32–39. doi:10.1016/j.geoderma.2008.05.002

Virtanen, S., A. Simojoki, H. Hartikainen, and M. Yli-Halla. 2014. Response of pore water Al, Fe and S concentrations to waterlogging in a boreal acid sulphate soil. Sci. Total Environ. 485–486:130–142. doi:10.1016/j.scitotenv.2014.03.071

Virtanen, S., A. Simojoki, O. Knuutila, and M. Yli-Halla. 2013. Monolithic lysimeters as tools to investigate processes in acid sulphate soil. Agric. Water Manage. 127:48–58. doi:10.1016/j.agwat.2013.05.013

Wallace, L.J., D.C. McPhail, S. Welch, D. Kirste, S. Beavis, S. Lamontagne, et al. 2006. Spatial heterogeneity of S and C stores in an inland acid sulfate soil. Geochim. Cosmochim. Acta 70:A683. doi:10.1016/j.gca.2006.06.1488

Webster, I.T., P.R. Teasdale, and N.J. Grigg. 1998. Theoretical and experimental analysis of peeper equilibration dynamics. Environ. Sci. Technol. 32:1727–1733. doi:10.1021/es970815g

Welch, S., D. Kirste, S. Beavis, G. Yates, F. Beavis, and L. Wallace. 2006. Geochemistry of sulfur in an inland acid sulfate soil system. Geochim. Cosmochim. Acta 70:A695. doi:10.1016/j.gca.2006.06.1512