Soil chemistry and acidification risk of acid sulfate soils on a temperate estuarine floodplain in southern Australia.
Acid sulfate soils (ASS) are soils or sediments that contain oxidisable or partly oxidised sulfide minerals (van Breemen 1988). Sulfidic material frequently underlies coastal and estuarine floodplains as well as mangrove swamps because these soils commonly form in waterlogged environments in the presence of labile organic material and abundant [Fe.sup.3+] and S[0.sub.4.sup.2-] (Andriesse and van Mensvoort 2006). Pyrite is the primary sulfide mineral that is formed (Dent 1986). Sulfuric acid is produced when sulfidic material is exposed to oxygen (Eqn 1), which commonly occurs on floodplains when the water table is lowered during periods of drought (Grealish et al. 2014; Mosley et al. 20146) or through artificial drainage of waterlogged land (Sammut et al. 1995, 1996; Mosley et al. 2014a). Acidity and other oxidation products can be discharged to adjacent waterways by both surface run-off and groundwater seepage, resulting in degradation of estuarine ecosystems (White et al. 1997; Johnston et al. 2004).
Fe[S.sub.2] + 15/4[O.sub.2] + 7/2[H.sub.2]O[right arrow] Fe[(OH).sub.3] + 2S[O.sup.2-.sub.4] + 4[H.sup.+] (1)
Our understanding of coastal ASS (CASS) in tropical and subtropical regions is well established (e.g. Wilson et al. 1999; Hicks et al. 2009; Johnston et al. 2009). ASS in eastern Australia have largely formed following the retreat of sea levels during the mid-Holocene highstand (Walker 1972; Wilson 2005). Sea level regression has been hypothesised to also be a driver of CASS formation in the northern Bass Strait region of southern Australia (Crawford and Rosengren 2012), yet knowledge of past highstands is poor, especially in this region (Lewis et al. 2013). Previous studies have characterised CASS formed in a Mediterranean climate (Poch et al. 2009), characterised and identified the effects of CASS oxidation during drought conditions (Fitzpatrick et al. 2010) and following reflooding (Sullivan et al. 2010) in the Lower Lakes region, where the Murray River discharges into the Southern Ocean. Although CASS have been mapped in some estuaries (e.g. Curdies and Moyne Rivers) and lakes (e.g. Lake Yambuk; Fitzpatrick et al. 2011), and we have some preliminary information about the presence of CASS in this region (Fitzpatrick et al. 2007), our knowledge of soil chemistry and distribution of CASS in temperate climates in northern Bass Strait is less well developed. In part, this is related to a lack of understanding of estuarine and associated floodplain evolution, which appears to be quite different in southern Australia compared with other eastern Australian estuaries (McSweeney et al. 2014). This means established models of CASS development (e.g. Lin and Melville 1993; Bourman et al. 2000; Wilson 2005) are unlikely to be directly applicable to the margins of Bass Strait.
Victorian estuaries, such as the Anglesea River, frequently suffer from low pH conditions during and after high-intensity rainfall events, which may often result in the closure of the estuary to recreational activities and subsequent downturn in the local economy (Corangamite Catchment Management Authority 2012). The acidification events have caused fish deaths in 2001, 2007 and 2010, with the most serious fish kill event occurring in 2001 (Smith et al. 2010). The fish kill events were attributed to the discharge of highly acidic river water and high concentrations of trace metals, such as Al, in the water column (Maher 2011). Several potential factors that caused the acidic river water and fish kill events have been proposed, including: (1) the export of acidity from acid mine drainage from an open-cut brown coal mine located in the upper catchment of the Anglesea River; (2) export of acidity from tea tree peat swamps, also in the upper catchment; and (3) the presence of CASS (Maher 2011). However, CASS in the Anglesea catchment have not been quantified and the extent of their contribution to the acidification events is unknown. The aim of this study was to characterise CASS and determine the potential acidification risk on the Anglesea River estuarine floodplain.
Materials and methods
Site descriptions and field sampling
The study site was located near the township of Anglesea, approximately 110 km south-west of Melbourne in southern Australia (Fig. 1). The region has an average annual rainfall of 630 mm, with the maximum rainfall occurring in August (69.2 mm) and minimum rainfall in March (34.8 mm). Average temperature varies between 23.0[degrees]C (February) and 7.3[degrees]C (July; Bureau of Meteorology (BOM) 2014).
The Anglesea River catchment is 885 ha with a river length of 20.5 km. The Anglesea River estuary is an intermittently closed and open estuary (McSweeney et al. 2014) which is 2.6 km long and 110 m wide near the mouth. The width of the estuary narrows to 40 m at 1 km upstream and to approximately 15 m in the uppermost reaches of the estuary.
Soil samples were taken from four sites: wetland (WE) and swamp scrub (SS) sites in the upper estuary, and woodland (WO) and coastal tussock saltmarsh (CTS) sites in the lower estuary on the Anglesea River floodplain (Fig. 1). Profiles were sampled using a hand auger to the water table and then with a gouge auger to a maximum depth of 1.30 m. Samples were placed into plastic polyethylene bags and air was excluded as much as possible. Soils were kept cool at 4[degrees]C until return to the laboratory and frozen until analysis commenced.
The WE site (38[degrees]24'06.3"S, 144[degrees]10'58.9"E; +1.72 m Australian height datum (AHD), where 0 m AHD is approximately sea level) was located in a flat area occasionally inundated with fresh to brackish water. Groundcover was dominated by samphire (Sarcocornia quinqueflora) and Austral seablite (Suaeda australis). Nine depths were sampled (0-0.10, 0.10-0.20, 0.20-0.30, 0.30-0.50, 0.50-0.70, 0.70-0.90, 0.90-1.10, 1.10-1.30 and 1.30-1.50m) and groundwater was at 1.3 m depth.
The SS site (38[degrees]23'52.9"S, 144[degrees]11'05.9"E; + 1.75 m AHD) was located in Coogoorah Park Nature Reserve. During extensive fires in the peat swamps in 1983, channels were excavated throughout the reserve to extinguish the fires, resulting in the formation of an island (Smith et al. 2010). The site was dominated by prickly tea-trees (Leptospermum continentale) and woolly tea-trees (Leptospermum lanigerum). Nine depths were sampled (0.10-0.20, 0.20-0.30, 0.30-0.50, 0.50-0.70, 0.70-0.80, 0.80-0.90, 0.90-1.10 and 1.10-1.30m). The groundwater table was at 0.35 m depth and a peat layer was found below 0.8 m depth.
The WO site (38[degrees]24'26.7"S, 144[degrees]11'26.5"E; +1.99 m AHD) was located approximately 250 m from the estuary channel and dominated by a range of Eucalyptus spp., including brown stringybark (Eucalyptus baxteri) and messmate stringybark (Eucalyptus obliqua). Seven soil depths were sampled (0.10-0.20, 0.20-0.30, 0.30-0.50, 0.50-0.70, 0.70-0.80, 0.80-0.90 and 0.90-1.10). The water table was at a depth of 0.2 m.
The CTS site (38[degrees]24'28.7"S, 144[degrees]11'22.9"E; +1.80 m AHD) was located approximately 150 away from the river and dominated by chaffy sawsedge (Gahnia filum) and prickly spear-grass (Austrostipa stipoides) with common occurrences of bearded glasswort (Sarcocornia quinquefloria). Five depths were sampled to 0.7 m (0.10-0.20, 0.20-0.30, 0.30-0.50 and 0.50-0.70). The water table was at 0.15 m depth.
Sample preparation and laboratory analysis
Samples were thawed under an [N.sub.2] atmosphere before analysis. Subsampled soils were oven dried at 85[degrees]C and crushed to pass through a 2-mm sieve. Soil pH, electrical conductivity (EC) and soluble cations and metals were determined on 1 :5 soil: water extracts (Rayment and Higginson 1992). Exchangeable cations and metals were extracted with 0.1 M Ba[Cl.sub.2]/N[H.sub.4]Cl extracts (Rayment and Higginson 1992). Soluble and exchangeable cations and metals were analysed by inductively coupled plasma mass spectrometry (ICP-MS) on a Perkin Elmer Elan DRC-e Inductively Coupled Mass Spectrometer.
Soil moisture content was determined by weight loss after oven drying at 105[degrees]C for 24 h. Oven-dried samples were crushed to pass through a 2-mm sieve and subsampled to determine particle size distribution. Oven-dried samples were then finely crushed to determine soil organic C (SOC), total N (TN), titratable actual acidity (TAA) and acid-neutralising capacity (ANC). The particle size distribution was measured by laser diffraction on a Beckham Coulter 13320 laser particle sizer after pretreatment with 5% tetrasodium pyrophosphate to remove organic matter. SOC and TN were determined by hightemperature combustion on an Elementar vario MICRO cube analyser. TAA was determined by titration with 0.05 M NaOH to pH 7.0 on 1:40 soil: 1 M K.C1 extracts after shaking for 4h (Rayment and Lyons 2011). ANC was determined by titration with 0.25 M NaOH to pH 7.0 on 1:25 soil:0.1M HC1 suspensions (Rayment and Lyons 2011). All samples were analysed in duplicate.
Reduced inorganic sulfur (RIS) was determined based on a three-step sequential extraction process (Burton et al. 2008). Acid volatile sulfide (AVS; which quantifies monosulfides) was extracted using the cold diffusion method of Hsieh et al. (2002) by shaking soils with 6 M HCl/0.1 M ascorbic acid. The evolved [H.sub.2]S was trapped in 3% Zn acetate in 1 M NaOH. Elemental sulfur ([S.sup.0]) was extracted with toluene by cold cyanolysis and [S.sup.0] was quantified spectrophotometrically according to the method described by Bartlett and Skoog (1954). Chromium reducible sulfur (CRS; which quantifies disulfides) was extracted by shaking with acidified Cr[Cl.sub.2] solution for 48 h. The evolved [H.sub.2]S was trapped in 3% Zn acetate in NaOH. AVS and CRS were quantified by iodometric titration with 0.025 M sodium thiosulfate. Samples were analysed in triplicate.
Net acidity was calculated by acid-base accounting, which was used to evaluate the potential for soil materials to generate acidity from sulfide oxidation and the neutralisation potential of the sediments (Ahem et al. 2004). Net acidity was calculated according to Eqn 2 and expressed in mol [H.sup.+] [kg.sup.-1].
Net acidity = (AVS + [S.sup.0] + CRS) + TAA - (ANC/FF) (2)
where a minimum fineness factor (FF) of 1.5 was used.
Soil properties of the sampled sites are shown in Fig. 2. Sites in the upper estuary (WE and SS) generally had a sandier texture (predominantly loams and sandy loams) compared with silt loams in the lower estuary (CTS and WO; Fig. 2a-c). In the upper estuary, the soil profiles were classified as a Sulfidic Redoxic Hydrosol at the WE site and a Sulfidic Oxyaquic Hydrosol at the SS site (Isbell 2002). The water table was located at a depth of 130 and 80 cm below the surface at the WE and SS sites respectively. In the lower estuary, the soil profiles were classified as a Sulfidic Extratidal Hydrosol at the CTS site and a Sulfidic Oxyaquic Hydrosol at the WO site (Isbell 2002). The water table, which fluctuates with tidal cycles, was located at a depth of 15 and 20 cm in the CTS and WO sites respectively.
The sites in the upper estuary had a relatively constant pH and EC throughout the profile. The soils in the WE site were moderately acidic (Fig. 2d) with lower EC relative to the other sites studied (Fig. 2e). In the SS site, the upper horizons (0-0.5 m) were highly acidic and with a lower EC, which increased below a depth of 0.8 m. In the lower estuary, the soils in the WO site were generally acidic with a high EC. In the CTS site, soils were alkaline with a high EC. The sites in the lower estuary generally had higher moisture content than the sites in the upper estuary (Fig. 2f).
TN concentrations decreased with depth from a mean ([+ or -] s.e. m.) 0.377 [+ or -] 0.020% in the 0-0.1 m layer to 0.011 [+ or -] 0.001% in the 1.3-1.5 m layer in the WE profile, and from 1.245 [+ or -] 0.033% to 0.003 [+ or -] 0.001% in the 0.5-0.7 m layer in the CTS profile (Fig. 2g). In the SS and WO profiles, TN decreased from 0.760 [+ or -] 0.010% and 1.577 [+ or -] 0.029% at the surface to 0.009 [+ or -] 0.002% and 0.503 [+ or -] 0.027% in the 0.5-0.7m layer, respectively, and then increased to 0.652 [+ or -] 0.022% and 1.382 [+ or -] 0.027%, respectively in the 0.9-1.1 m layer.
The WE site had relatively low SOC concentrations, ranging from 2.125 [+ or -] 0.085% to 0.504 [+ or -] 0.009%. Conversely, the CTS had much higher SOC concentrations, ranging from 14.583 [+ or -] 0.420% to 0.358 [+ or -] 0.016% (Fig. 2h). The SS profile had higher SOC concentrations in the 0-0.1m layer (16.744 [+ or -] 0.068%) which decreased (to 0.603 [+ or -] 0.030%) in the 0.2-0.3 m layer and then increased (to 21.940 [+ or -] 0.567%) at the 1.1-1.3 m depth. The SOC concentrations in the WO profile exhibited a similar pattern to that seen in the SS profile. The surface layers in the WO profile had a maximum SOC concentration of 26.716 [+ or -] 0.496% at the surface, which decreased to 1.109 [+ or -] 0.069% in the 0.2-0.3 m layer and increased to 22.452 [+ or -] 0.457% in the 0.7-0.9 m layer before decreasing further with depth.
Soluble and exchangeable cations and metals
The concentration of soluble S[O.sub.4.sup.2-] in each of the sampled profiles is shown in Fig. 3a. The concentration of soluble S[O.sub.4.sup.2-] throughout the profile in all sites exceeded the trigger value (1.04 mmol [kg.sup.-1]) for potential formation of monosulfide (Bush et al. 2009), with the exception of the WE site below 1.2 m depth. The [Cl.sup.-]/S[O.sub.4.sup.2-] ratios of all soil extracts in the SS, WO and CTS sites and in the upper horizons in the WE site were less than 4 (Fig. 3b).
The [Cl.sup.-]/[Na.sup.+] ratios in all soil extracts in the SS, WO and CTS sites were less than that typically found in seawater (Fig. 3c). The upper horizons in the WE site had a greater [Cl.sup.-]/[Na.sup.+] ratio and smaller [Na.sup.+]/[Mg.sup.2+] ratio relative to seawater.
The release of [K.sup.+] occurred in the SS site, which was indicated by the [Ca.sup.2+]/[K.sup.+] and [Na.sup.+]/[K.sup.+] ratios (Fig. 3c, f). The [Na.sup.+]/[K.sup.+] ratios in the SS, WO and CTS sites were greater than that of seawater. However, the [Ca.sup.2+]/[K.sup.+] ratio in the SS site in the upper estuary was smaller than in seawater.
The [Na.sup.+]/[Ca.sup.2+] ratios in the horizons in all sites, except the 0.1-0.8 m layer of the SS profile, were less than that for seawater (Fig. 3g).
Concentrations of soluble metals are shown in Fig. 3h-k. In general, the concentration of soluble Fe was lowest at the surface in the WO site (Fig. 3h) in the lower estuary. Al concentrations were high in both the surface layers and at depth below 0.8 m at all sites (Fig. 31). The concentrations of Mn and Zn were also high in the WO profile (Fig. 3j-k).
The concentrations of exchangeable cations and metals in the soil extracts are shown in Fig. 4. In general, the upper horizons in the WO and CTS sites had relatively high concentrations of exchangeable base cations ([Ca.sup.2+], [Mg.sup.2+], K and [Na.sup.+]; Fig. 4 a-d). The concentrations of exchangeable [Fe.sup.2+]/[3.sup.+] and [Al.sup.3+] were very high in the surface layers of the SS site (maximum 8.12 and 33.4[cmol.sub.c] [kg.sup.-1] respectively). Conversely, the highest exchangeable A1 concentrations occurred in the WO site at depth (maximum 88.5 [cmol.sub.c] [kg.sup.-1]). Concentrations of exchangeable [Mn.sup.2+]/[3.sup.+] and [Zn.sup.2+] were also highest in the WO site.
Acid generation and neutralisation potential
AVS was present in most of the horizons at sites in the upper estuary (Fig. 5a, b) with a maximum concentration of 0.05% S in the WE site in the 0.7-0.9 m layer. Flowever, AVS was only detected in one horizon in the WO site and two horizons in the CTS site (Fig. 5c, d). Elemental [S.sup.0] was not detected at any site (data not shown). CRS was only detected in the sites in the upper estuary, with a maximum concentration of 0.03% S in the 0.2-0.3 m layer at the WE site.
The TAA of each site is given in Fig. 3e. In the upper estuary, all the horizons in the WE and SS sites, except the surface soil in the WE site, had a positive TAA value. In the SS site, the high TAA occurred at the surface (203.4 mol [H.sup.+] [t.sup.-l]) and at depth (205.6 mol [H.sup.+] [t.sup.-l]). In the lower estuary, the CTS site had a negative TAA throughout the depth, whereas the WO site had a high positive TAA in the lower horizons, with a maximum of 253.6 mol [H.sup.+] [t.sup.-l].
The ANC of the sampled sites is shown in Fig. 5f The sites in the upper estuary generally had a lower ANC than sites in the lower estuary. ANC was not detected in four horizons, including at the surface in the SS site.
The WE and CTS sites had a negative net acidity throughout the profile (Fig. 5g). In the SS site, the surface layer had a very high positive net acidity (i.e. >100.0 mol [H.sup.+] [t.sup.-l]) and the lower horizons had a moderate to high positive net acidity (i.e. >19.0mol [H.sup.+] [t.sup.-l]). The lower horizons in the WO site had a moderate to high positive net acidity.
Potential acidification risk
The present study identified potential sources of acidity and characterised the soil chemistry of CASS present on the estuarine floodplain. The potential acidification risk to the floodplain and estuary increases with distance upstream largely due to limited buffering capacity and higher contributions of R1S in the sites located further upstream.
The accumulation of A VS at the surface and at depth in the upper estuarine floodplain can pose a significant hazard to the environment (Fig. 5a, b). In most estuarine systems, the dominant RIS species is generally pyrite (Fe[S.sub.2]), whereas monosulfides and [S.sup.0] typically comprise a minor fraction of the total RIS (Reddy and DeLaune 2008). However, these results indicate that monosulfides are the dominant form of RIS in the Anglesea River estuary floodplain, which tends to increase in concentration further upstream (WE and SS; Fig. 5a-d). AVS is metastable and highly reactive because oxidation of monosulfides can occur much more rapidly (i.e. hours) than the oxidation of disulfide minerals such as pyrite (i.e. days; Burton et al. 2009). Sulfidic material in the WE and SS sites close to the surface can be exposed to oxidising conditions when water table levels are lower, such as when sampling occurred in the present study. Oxidation products in the subsoil can be moved upwards to the surface by capillary action and evapotranspiration processes during periods of low water levels (Lin et al. 1995). This is potentially reflected in the high concentrations of dissolved Fe and Al in the surface soils in the upper estuary (Fig. 3h, i). These metals can potentially be mobilised into adjacent waterways either by surface run-off or through the banks of the channels by lateral flow. Lateral seepage occurs when the groundwater level is between the surface and the minimum low tide level in the adjacent waterway resulting in high concentrations of trace metals leaching into the river (Johnston et al. 2004). Therefore, a prolonged dry period, when water levels remain low for an extended period of time, can potentially lead to an accumulation of acidic oxidation products that can be rapidly transported to the adjacent channels and cause a rapid decrease in surface water pH (e.g. Sammut et al. 1996; Wilson et al. 1999; Mosley et al. 2014a, 20146). Relatively lower concentrations of soluble metals at the SS site in the upper portion (0-0.6 m) of the soil profile at low pH values may be due to vertical and lateral leaching as a result of the sandier texture.
In intermittently closed and open lagoons such as the Anglesea River estuary, stratification of the estuarine water body can potentially occur when the estuary mouth remains closed or partially closed for an extended period of time (McSweeney et al. 2014). Stratification of the water column results in clear differences in temperature, salinity, pH and dissolved oxygen. After heavy rainfall events, rapid declines in surface water pH due to transport of acidity and associated trace metals are likely to contribute to large fish kill events in the estuary. Additional sources of acidity may also be located in peat swamps located in the upper catchment, which is transported to the estuary following heavy rainfall events (Maher 2011). Localised precipitation of transported metals, such as Al, can also occur during periods of stratification when high concentrations are mobilised due to pH buffering at the halocline (Pope 2006). However, further sampling of both soil and water and monitoring of water quality throughout the year is required to identify the multiple sources of acidify, including sources from groundwater and surface water.
High concentrations of RIS in the sites located on the upper estuarine floodplain (WE and SS) and limited formation of RIS in the sites located on the lower estuarine floodplain (CTS and WO) may be related to small variations in elevation, with WE and SS located at a lower elevation than CTS and WO (Fig. 1b). Therefore, the WE and SS sites may be subjected to longer periods of inundation, which encourages formation of RIS. However, the quantity of CRS formation in the WE site is relatively high compared with the SS site. This may be due to frequent prolonged waterlogged conditions at the WE site, which allow the transformation of monosulfides to the more stable form of pyrite (Keene et al. 2011). The results show that the dissolved S[O.sub.4.sup.2-] is not a limiting factor for the formation of RIS because its concentration in both sites in the low estuary exceeds the trigger value (>1.04 mmol [kg.sup.-1]).
SOC concentrations are generally highest at the surface at all sites. The large increase in soil carbon concentrations in the layers below 0.7 m depth at the SS and WO sites (Fig. 2h) coincides with the water table resulting in reducing conditions. When water tables are high, oxygen availability is limited, which slows the decomposition of organic matter and leads to an accumulation of SOC deeper in the profile (Ponnamperuma 1972). The presence of SOC concentrations in the SS and WO profile also coincides with lower pH (Fig. 2d) and high TAA (Fig. 5). Root respiration and decomposition can increase the partial pressure of C[O.sub.2] and contribute to slight decreases in pH in the soil pore water. In addition, organic acids produced from an excretion of protons from plant roots may also result in a decrease in pH, whereas carboxyl groups on organic matter can also contribute to acidity. We suggest that the high organic matter content, reflected in SOC concentration, can potentially contribute to TAA concentrations in the SS and WO sites. Below the 0.8 m depth layer in the SS site, the TAA is up to 205.6 mol [H.sup.+][t.sup.-1], which is up to fourfold higher than the TAA higher in the profile (Fig. 5e). Similarly, the TAA concentration is up to 277.8 mol [H.sup.+][t.sup.-1] below the depth of 0.4 m at the WO site.
The [Cl.sup.-]/S[O.sup.4.sub.2-] ratios of all the soil extracts in the SS, WO and CTS sites and the upper horizons in the WE site were less than 4 (Fig. 3b), suggesting additional sources of S[O.sub.4.sup.2] which may have originated from sulfide oxidation (e.g. Mulvey 1993). The [Cl.sup.-]/[Na.sup.+] ratios at the SS, WO and CTS sites were less than the ratio typically found in seawater (Fig. 3c), suggesting that Na+ from seawater may have been sorbed on the layer silicate and Fe/Al oxide minerals. This process could also be reflected in the greater [Na.sup.+]/[Mg.sup.2+] ratio compared with seawater, especially in the SS and CTS sites (Fig. 3d). The [Na.sup.+]/[Ca.sup.2+] ratios at all sites, except the 0.1-0.8 m layer of the SS profile, were less than that for seawater (Fig. 3g), suggesting sorption of [Na.sup.+] and acid-induced dissolution of shell fragments, as discussed below (Sammut et al. 1996).
The higher EC and pH in the sites located on the lower estuarine floodplain (Fig. 2e) reflect the influence of seawater (Fig. 2d). Seawater is a source of alkalinity, which has up to 2.5 x [10.sup.-3] M of alkalinity, predominantly in the form of HC[O.sub.3.sup.-] and HC[O.sub.3.sup.-] (Stumm and Morgan 1996). Therefore, seawater contributes a high ANC to the soils, reflected in the WE and CTS sites (Fig. 5f), which leads to the high negative net acidify throughout the profiles at those sites (Fig. 5g). In addition to seawater, Ca[CO.sub.3] shell material also contributes to ANC in the lower estuary. In the CTS site, carbonate materials were identified in the sand layer at 0.7 depth following field testing and reaction with HC1. The carbonate materials contribute a high ANC at the 0.5-0.7 m depth that is fourfold higher than the ANC of the overlying horizons (0.2-0.5 m depth). The low [Cl.sup.-]/S[O.sub.4.sup.2-] values indicate that sulfidic materials at the sites have been oxidised in the past (Fig. 3b; Lin et al. 1998) with the acid generated from the sulfide oxidation neutralised by shell fragments, because the [Na.sup.+]/[Ca.sup.2+] ratios in the sites are smaller than that for seawater (Fig. 3g). Therefore, although A VS is present, oxidation of RIS is unlikely to result in acidification because of the high ANC in the lower estuary.
Conversely, the influence of seawater is limited in the upper estuary and with increasing distance from the main estuarine channel, with brackish water contributing less alkalinity to the soils in the upper estuarine floodplain, consistent with the EC and pH values (Fig. 2e). In the SS and WO sites, the ANC in most of the horizons was low or undetected (Fig. 5f). Increasing distance upstream and from the main channel constrains the distribution of alkalinity from seawater. Shell was not found at the SS and WO sites, resulting in high positive net acidity (Fig. 5g).
Artificial estuary openings
The Anglesea River estuary is an intermittently open and closed estuary, characteristic of this region of southern Australia (McSweeney et al. 2014). When the estuary mouth is closed due to longshore accumulation of sand, the estuary is dominated by freshwater inputs, whereas marine inputs dominate when the estuary mouth is open (Pope 2006). The estuary mouth can be opened naturally following high rainfall events in the upper catchment or, more recently, via artificial openings. Artificially opening the estuary mouth, such as what occurred in 2010 and 2011, can rapidly introduce and inundate the upper estuarine floodplain with brackish water during high tides and lower water levels during low tides (Arrowsmith et al. 2010).
Although the [Cl.sup.-]/[Na.sup.+] ratio in the sites in the lower estuary is smaller than seawater, the ratio in the SS site is also unexpectedly low (Fig. 3c). The high EC and low [Cl.sup.-]/[Na.sup.+] ratio suggest that deposition of marine salts has occurred in the past. Wong et al. (2010, 2015) found that short-term inundation with seawater was associated with the displacement of trace metals due to increasing ionic strength of the solution. A significant finding in Wong et al. (2010) was that even brackish water with salt concentrations of 10-50% of seawater concentrations can desorb high concentrations of trace metals and mobilise acidify. This can potentially contribute to acidic discharges when the estuary mouth is artificially opened. The displacement of acidic metals can also be significant in the SS and WO sites because there are large exchangeable acidic Al, Fe and Zn pools available for mobilisation (Fig. 4e, f h). The effects of estuary mouth opening on CASS in this region warrant further research.
It is possible that CASS in the northern region of Bass Strait in southern Australia are different to the CASS previously studied and mapped on the east coast of Australia (e.g. Walker 1972; Lin and Melville 1993; Wilson 2005). These differences may be related to the age of formation, the high ANC and geomorphic influences, such as the narrow bedrock-controlled floodplain, in which the CASS form. It is also possible that the coal seams of the Otway Basin, which extend to the west of the study area (Holdgate 2003), also contribute to the differences seen in this region. However, this environmental setting is not unique to the Anglesea River catchment. Extensive natural coal seams are also common in the Gippsland Basin, located to the east of the study area, and which are found upstream of the coastal plain where CASS have been identified (Fitzpatrick et al. 2011). These coal seams may also act as a source of metals in the sediments. Similarly, the intermittent opening and closing of the Anglesea River estuary mouth, which is characteristic of several estuaries in Victoria (McSweeney et al. 2014), will also affect the geomorphic processes related to CASS formation. Approximately 90% of estuaries in Victoria intermittently open and close, including the majority of those assessed in Fitzpatrick et al. (2007), and those located both to the east and west of the study area.
It is possible that CASS may extend further inland than 1 m above the current mean sea level, because sea levels may have been higher than previously assumed during the mid-Holocene (Wilson 2005; Lewis et al. 2013). If this is the case, then this may have implications for potential acidification of the Anglesea estuary with limited buffering capacity evident in the upper estuary sites that were examined. This also requires further investigation.
CASS were identified on the Anglesea River estuarine floodplain; however, the potential for acidification varied. In the lower estuary, the potential acid generated from the oxidation of sulfidic sediments can be neutralised by shell materials and seawater, resulting in negative acid-generation potential. The effects of seawater and shell material diminish upstream and with increasing distance from the estuarine channel. Therefore, in the upper estuary where there was limited ANC, acid-generation potential was positive. High concentrations of SOC in the topsoils and the layers below 0.7 m depth of the SS and WO profiles further contribute to acidity via the contribution to TAA. Consequently, the SS and WO sites are likely to acidify soils and, potentially, the estuary. High concentrations of trace metals can further contribute to acidification when pH decreases, particularly at depth in sites located on the upper estuarine floodplain.
Received 19 June 2015, accepted 14 January 2016, published online 22 August 2016
The authors acknowledge Ursula Pietrzak-Aniszewska for assistance in the laboratory and the Environmental Analysis Laboratory, Southern Cross University for sample analysis. This research was funded by the Corangamite Catchment Management Authority.
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C. C. Yau (A), V.N.L. Wong (A,C), and D.M. Kennedy (B)
(A) School of Earth, Atmosphere and Environment, Monash University, Wellington Road, Clayton, Vic. 3800, Australia.
(B) School of Geography, The University of Melbourne, Vic. 3010, Australia.
(C) Corresponding author. Email: email@example.com
Please note: Some tables or figures were omitted from this article.
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|Author:||Yau, C.C.; Wong, V.N.L.; Kennedy, D.M.|
|Date:||Oct 1, 2016|
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