Mechanisms of acid sulfate soil oxidation and leaching under sugarcane cropping.
The acid sulfate soils (ASS) on the New South Wales north coast, Australia, are located in what were originally tidally influenced floodplain areas, adjacent to ecologically sensitive land and aquatic ecosystems. The large tidal range, small catchment size, and low outflows characteristic of eastern Australian embayments have resulted in the extensive deposition of typically fine-grained Holocene sulfidic sediments (White et al. 1997). Indeed, concentrations of pyrite have been previously measured at > 3% over most of the ASS profile within the McLeod's Creek catchment (van Oploo 2000), a tributary of the Tweed River. The oxidation of these sediments and the export of the oxidation products (particularly acidity and dissolved metals) has resulted in significant environmental impacts, including massive fish kills and outbreaks of epizootic ulcerative syndrome, within the study area (Easton 1989), as well as in other eastern Australian estuaries (Callinan et al. 1993; Sammut et al. 1993, Sammut et al. 1995, 1996; Willett et al. 1993).
Most of the Tweed River floodplain is used for broad-acre sugarcane production with well-established drainage systems. It is commonly believed that drainage systems with 1-way flap gates have caused the observed oxidation of pyritic sediments and the associated discharge of acidity by the lowering of the watertable relative to the sulfidic sediments (Walker 1972). However, Lin et al. (1995) demonstrated that oxidation of ASS landscapes occurred even without drainage.
Smith et al. (2003) showed that the concentration of existing acidity stored in ASS profiles is in the order of 50 t [H.sub.2]S[O.sub.4] equivalent/ha, but only about 100-200 kg [H.sub.2]S[O.sub.4]/ha is exported from the Tweed floodplain each year (Wilson et al. 1999). Greatest concentrations of existing acidity occur along original natural drainage lines (Smith et al. 2003). Understanding the sources and mechanisms of transfer of this acidity is extremely important to the health of downstream aquatic ecosystems. Any insight regarding the transfer of the chemical constituents within the ASS profiles under sugarcane also allows for the development of appropriate management strategies. The extent of drainage-induced oxidation and its effects on the transfer of oxidation products will therefore be extremely site-specific, depending on the geomorphology of the area, the antecedent hydrology and drainage system in place, and the localised watertable fluctuations.
Therefore, the specific aim of this paper is to examine the concentrations of soluble ions at a representative location on the Tweed River, and by inferring the transport processes occurring, suggest possible management/amelioration strategies that may minimise the transfer of such oxidation products from the system. It should be noted that this is a preliminary study, and further measurements need to be made to confirm the trends outlined, as well as to identify whether the results are comparable to other areas along the Australian coastline. The issue of anthropogenic v. natural causes of ASS oxidation and acidity production will also be addressed.
Materials and methods
The soil sampling was undertaken during February 2000 on the backswamp floodplain of McLeod's Creek (28[degrees]18'S, 153[degrees]31'E), a tributary of the Tweed River, north-eastern NSW (Fig. 1). The climate of the Tweed River is humid and subtropical, with an average annual rainfall of > 1400 mm, and a pronounced wet season extending from December to March (Wilson et al. 1999). McLeod's Creek was first modified in the 1930s to remove water from wetland areas for cattle pasture and later for sugarcane production (van Oploo 2000). It has subsequently been widened and straightened, and with the construction of field drains approximately 30 years ago, there is currently > 100 km of drains with McLeod's Creek being the largest (Wilson et al. 1999).
Sixteen soil profiles were extracted from a single canefield in 2 parallel transects crossing over a typical field drain. The profiles were taken at intervals of 0, 0.5, 1, and 5 m perpendicular to, and on each side of, the drain. One side of the drain was fallow, the other under cane. Of the soil profiles extracted and described in the field, 6 were chemically analysed in the laboratory, with the remainder used for watertable measurements. The configuration of the analysed profiles is shown in Fig. 2. From the field description, which included pH, reaction to peroxide, and physical attributes, layers within the profiles were identified and representative samples were selected for laboratory analysis. These included material from the topsoil, the jarosite layer, above and below the oxidation front, and from the deep, unoxidised blue-grey gel. The selected samples were sealed in airtight plastic bags and frozen for transportation back to the laboratory. At each profile, a 0.1-m-diameter PVC pipe was inserted into the space where the soil had been removed, enabling the subsequent measurement of watertable elevations at regular intervals. The ground surface was accurately surveyed, relative to Australian Height Datum (AHD).
Standard laboratory methods were used to measure the concentrations of soluble ions and acidity in ASS profiles at 0.1-m depth increments in the 6 profiles. The extracted samples were rapidly dried in an oven at 85[degrees]C for 24 h and allowed to cool. Laboratory analysis included 1 : 5 soil : deionised water extracts measuring a range of soluble cations and anions by ICP-OES (aluminium, Al; calcium, Ca; iron, Fe; potassium, K; magnesium, Mg; manganese, Mn; sodium, Na; sulfur/sulfate, S[O.sub.4]; silicon, Si; strontium, Sr). Further soil : deionised water extracts were used to measure electrical conductivity, chloride ion (CI) concentration (using an Activon chloride combination ion electrode), and acidity by titration to pH 5.5 with sodium hydroxide. Total actual acidity (TAA) was measured by titration with sodium hydroxide to pH 5.5 on 1 : 5 soil : 1 M KC1 extracts. Total potential acidity (TPA) was also conducted on 1 : 5 soil : 1 M KC1 extracts treated with hydrogen peroxide. Both TAA and TPA measurements were performed in accordance with the analytical methods set out in the Acid Sulfate Soils Management Advisory Committee (ASSMAC) Guidelines (Stone et al. 1998).
The field descriptions show an organic surface layer characterised by a brown to dark brown colour, and richness in organic matter. It extends to a depth of around 0.0 m to -0.2 m AHD, and its pH is generally >4.0. The oxidised sulfuric layer is characterised by a distinct colour change to lighter brown, along with mottling of yellow jarosite, and red/brown iron oxides/hydroxides, possibly goethite. The B layer extends to a maximum depth of -0.5 m AHD, and its pH ranges from 2.7 to 4.0. The transition zone exhibits qualifies of the actual ASS above it, and the potential ASS below it. There is a gradual change in texture from a brittle or crumbly consistency to a labile or plastic texture towards the bottom of the transition zone. The pH also traversed from acidic at the top (~3.5) to near neutral at the bottom (~6.0). Planes of iron oxidation and jarosite still occur in this layer, along with fine roots. The sulfidic or potential ASS layer consists entirely of unconsolidated blue-grey gel, extending beyond the depth of sampling (-1.15 m AHD). The pH slowly increases down the profile from near-neutral to neutral (~6.9). Intermittent depths contain fine roots and shell fragments. No jarosite or iron oxidation products occur at these depths.
A fundamental requirement for the use of any water quality or soil extract dataset is that the total charge of anions should equal the charge of cations. Here, the charge-sum of water soluble anions (y) approximately equals the sum of water-soluble cations (x) ([R.sup.2] = 0.99,y = 1.06x -8.52, n = 33). In addition, there is a clear linear relationship between electrical conductivity (y) and the sum of water soluble cations (x) ([R.sup.2] = 0.97,y = 0.02x - 0.02, n = 33). The latter 2 variables are measured independently of one another. The high correlation coefficients for these 2 relationships therefore allow confidence in making subsequent interpretations of the presented data.
All soluble species were shown to have smaller concentrations in the topsoil and sulfuric layer than the underlying sulfidic sediments. Several distinct trends in the soluble ion chemistry were observed along the transect, when moving away from the field drain edge towards the centre of the sampled canefield (Fig. 3). Most noticeable was the decrease in C1 and Na, simultaneously coupled with the increase in S[O.sub.4], Ca, and Mg when moving from the drain edge (0 m) to within the canefield (5 m).
The acidic cations (Al, Fe, and Mn), which are shown in Fig. 4, have elevated concentrations in the 5-m profiles compared to the 0-m profiles. Within both the 5- and 1-m profiles, the concentration of each ion reaches a maximum at the top of the transition zone. This fact is masked within the 0-m profile with the reduced values of all ions.
The depth-weighted average, calculated for a single transect, showed the TAA was >2.5 times greater within the canefield (5 m) than at either 1 or 0 m from the field drain edge. Values ranged from 46.14 t [H.sub.2]S[O.sub.4]/ha at the 5-m profile to 19.97 and 18.92 t [H.sub.2]S[O.sub.4]/ha, at the 1- and 0-m profiles, respectively. As these calculations were performed on only a single transect, further repetition of the analysis is needed to confirm this trend. However, the values are commensurate with the 45 profile measurements made by Smith et al. (2003) for profiles within cane blocks at the study site.
Comparison of the TAA and TPA results (Fig. 5) shows the extensive store of potential acidity throughout the solum, with a large increase below the oxidation front, where the pyrite remains unoxidised. The elevated levels of potential acidity in the sulfidic layer (~450 mol H+/t) compared with the TAA at the same depth (<20 mol H+/t) represents the actual ASS overlying the potential ASS, typical of any ASS environment. The elevated levels of TPA compared with TAA in the sampled topsoil can only be explained by the limitations of the TPA methodology, particularly the interference of minor amounts of sulfate minerals and organic matter (Clark et al. 1996; Lin et al. 1996; Sullivan et al. 1999). More importantly though, the TPA results in Fig. 5 highlight the relatively uniform concentration of potential acidity along the transects. Therefore, the data collected are consistent with the theory that most acidity from protons and hydrolysable ions released into the adjacent drain water is coming from only within 1 m of the field drain edge.
When the field descriptions of the soil profiles were compared with watertable measurements taken from the extracted profiles, it was apparent that at the time of sampling the watertable existed primarily within the sulfuric layer, as shown in Fig. 6. The observable increase in elevation of all layers at 0.5 m from the field drain is due to the compaction of the ground by tractors and other heavy machinery on the adjacent cane-growing surfaces, a point reinforced by the vertical exaggeration of the figure. It should be noted that Fig. 6 represents the watertable during the February soil sampling period, which is during the middle of the wet season for the area. Rainfall measurements preceding and during the time of sampling were less than the average annual measurements previously recorded. Similar measurements taken during April and July of that year showed the same outcome, that the watertable existed above the oxidation front.
The results of the field descriptions and detailed laboratory analysis suggest that under the current drainage regime, most of the acidity and other soluble ions generated from pyrite oxidation and being exported into the drain is doing so <5 m from the field drain edge. Leaching and mass movement associated with watertable fluctuations and local rainfall are primarily responsible for the removal of ionic species at the field drain edge, and within the upper layers of the soil profile. Solute movement within the unsaturated zone will also be influenced by the nutrient uptake of growing sugarcane and by adsorption and exchange within the soil's exchange complex, as well as by convective and dispersive forces that will be especially active close to the field drain edge where concentration gradients are at their maximum. These points are illustrated by the progressive decrease in acidity, sulfate, and higher valence metal ions away from the centre of the canefield, dispersed from the point of generation. The converse can be seen with the Na and C1 profiles, which show a progressive decrease into the canefield, away from their actual source, the brackish drain water.
These effects decrease slightly within the lower layers, due in part to the rapid decrease in hydraulic conductivity measured at depth (White et al. 1993), and also due to locations further into the canefield. The low pH and large concentrations of sulfate in the upper horizons indicate that capillary action and diffusion play a vital role in the transfer of oxidation products within the sampled canefield (i.e. > 1 m away from the field drain edge). It is likely that these mechanisms transfer most of the oxidation products upwards along moisture potential gradients during low watertable regimes under such artificial drainage conditions. This interpretation is consistent with that identified by Lin et al. (2001), also at McLeod's Creek. Downward leaching of dissolved ions during rain-fed infiltration is probably negated by the upward movement of the watertable in response to the infiltration. The overall affect of rainfall infiltration, upward watertable and capillary concentration of salts, is for a net leaching regime of upward-and-out to surface drainage (Fig. 7).
Natural v. anthropogenic oxidation of ASS landscapes
The Australian ASS literature appears to generally propose that European drainage of backswamp ASS landscapes has been the predominant cause of the observed oxidation and acidification problems. We have been a party to this view (e.g. see Callinan et al. 1993; White et al. 1993; Willett et al. 1993), perhaps greatly influenced by the earlier literature based on drainage-induced acidification in Dutch polders that were previously shallow coastal seas (e.g. Pons 1973). This is also unambiguously the case in Australia where bund walls and flap-gates were constructed in permanently tidal-saturated landscapes, such as in the mangrove swamps on the southern shore of Trutes Bay (a left-bank tidal embayment of the lower Tweed River), or on the right-bank shore of East Trinity Inlet, Queensland (see Cook et al. 2000). However, in estuarine floodplain backswamps, the belief of a European drainage-caused landscape oxidation appears to be founded on the general observation that when artificial drainage systems were first installed, increased acidic discharge and their environmental impacts occurred. However, such observed acidic discharge would also occur if the drainage system only provided the new conduit for enhanced acidity export from an already naturally oxidised landscape. We now propose the importance of this latter, natural pedogenesis, acidity cause, based on many field observations (e.g. Lin et al. 1995; Donner and Melville 2002; Smith et al. 2003), and this has been accepted by the NSW Healthy Rivers Commission (HRC 2003, pp. 47-48). Unfortunately, it is now difficult to test natural pedogenesis satisfactorily because most of these floodplain backswamps on east coast Australia have already had some degree of drainage, and no sufficiently detailed pre-drainage analysis of ASS and the landscape hydrology has been undertaken.
It is logical that such drainage systems markedly reduced the duration of backswamp inundation (White et al. 1993), but it has not been established that the changed degree and duration of backswamp inundation altered watertable elevations and ASS profile saturation sufficiently to cause the observed, almost ubiquitous, > 1 m depth of ASS profile oxidation. The lowering of the watertable should be seen as preferential oxidation down drained, generally vertical macropores. To some degree this does occur; however, most commonly at McLeod's Creek and at other sites where such drainage systems exist, a uniformly horizontal oxidation front occurs within the saturated sediment and below any observable watertable elevation (Wilson et al. 1999; van Oploo 2000). This oxidation depth of generally >1 m exists in the fine-grained sulfuric/sulfidic sediment with a very small pore size that gives an effective capillary fringe and saturated sediment near to the ASS mineral profile's surface. To drain the pores in this sediment would require a watertable elevation much lower (> > 1 m) than any observed in the past 15 years of study at McLeod's Creek. This period recently coincided with the most severe and prolonged drought recorded in European settlement of the Tweed area. Such an apparently uniform and deep oxidation front is also observed where no artificial drainage has occurred (see Lin et al. 1995). The existence of this oxidation front below the watertable appears most likely due to downward diffusion of some oxidants through the saturated soil. This is currently under research at McLeod's Creek so that a clearer picture of ASS landscape pedogenesis may be obtained.
Some authors (Blunden and Indraratna 2000; Johnston et al. 2002) have presented results where at least some of their study site materials are sandier and/or apparently have a greater degree of lateral hydraulic conductivity than that described for Mcleod's Creek. Nevertheless, these represent exceptions to the nature of the materials deposited in the geomorphic mud basins that generally formed behind east coast sand barriers.
Coastal geomorphology, sea level, and climate changes during the Holocene
The geomorphological variations in sediment characteristics and stratigraphy are not well recognised in ASS research, so that the oxidised sediment at the top of the ASS profile is presumed to have been the same as that below the oxidation front, particularly with respect to the original sulfide mineral content. This is unlikely to be the case because, whereas the deeper estuarine sediment was deposited in brackish tidal water conditions (rich in dissolved sulfate), the material closer to the surface will have had an increasing fluvial influence. Walker (1970) provided an excellent early view on geomorphic/pedogenic development in the estuarine/fluvial evolution of the lower Macleay River floodplain. Dalrymple et al. (1992) and Roy et al. (2001) also provide models to explain the infilling of east coast estuarine embayments with their sand-barrier-induced, inner mud basins where sulfidic sediments accumulated initially, later to be overtopped by fluviatilc sediment. The degree of estuary embayment infilling and emergence of the mud basin surface above a shallow tidal brackish lake will vary, depending on sediment input rate and the relative sea level in the estuary embayment. There are some estuary embayments that are 'mature' and completely infilled (e.g. most of the Clarence embayment) and other 'immature' embayments where significant parts have not been infilled (e.g. Cudgen Lake).
The history of the last global post-glacial sea level rise from about -125 m at 20 ka to its present position at about 6.5 ka is now well established (see Lambeck and Chappell 2001). Thom and Roy (1985) proposed their models of east coast Australian geomorphic evolution based on a constant sea level since about 6.5 ka. This is likely to be true of the sea level as controlled by the balanced inputs/outputs of a constant water volume into a globally constant ocean basin volume. However, recent research on east coast Australia shows that sea level reached a maximum at about 6.5 ka but this was up to 1-3 m above today's relative sea level. Isostatic readjustment of the lithosphere along the Australian continental margin caused coastal uplift, but the degree of this uplift varied even at locations only 100 km apart, and increased with distance inland from the coast (Lambeck and Nakada 1990). Some work suggests this uplift rate (and effective sea level fall) has been constant until the present (see Nakada and Lambeck 1989; Lambeck and Nakada 1990; Lambeck and Chappell 2001). However, Flood and Frankel (1989), Baker and Haworth (1997), and Baker et al. (2001) used fossil intertidal tube worm deposits stranded on coastal cliffs to infer a more prolonged, constant high relative sea level until about 2 ka, then a fall to present relative sea level. Whatever the timing of a relative sea level fall by up to 1-3 m, the magnitude and duration of this drainage base level decline is sufficient to account for much of the observed backswamp ASS oxidation, certainly where contact with tidal influence has been maintained.
The weather patterns of today are unlikely to be representative of the climate throughout the Holocene epoch (< 10 ka), as can be seen in the first instance from variations in rainfall and associated flooding during the second half of European settlement. Pittock (1975) identified a 10 -20% increase in mean annual rainfall over much of NSW and Queensland during the 1940s to 1970, compared with the first half of the 20th Century. W. D. Erskine and his colleagues extended this study to include records from the 19th Century and later 20th Century and proposed multi-decadal periods (about 50 years) of drought-dominated rainfall regime (about 1890-1946) and flood-dominated rainfall regime (< 1890 and 1946+), mostly with increased summer rainfall. Bell and Erskine (1981), Erskine (1986), and Erskine and Warner (1988) showed this latter increase in rainfall in the Nepean and Hunter Valleys gave an upward shift by 50-100% in the flood frequency curve. Smith and Greenway (1983) also showed an increase in flood height on the Tweed, Richmond, and Clarence Rivers after the mid 1940s. We do not have evidence of the existence of these weather patterns before European settlement, but the global 'Medieval Warm Period' (about 1100-1300) possibly had mean temperatures several degrees Celsius warmer than today. Elevated temperatures will be expressed in increased evapotranspiration, which has been shown as a major control on floodplain watertable elevation (e.g. White et al. 1997).
The south-eastern Australian climate of the Holocene, since present sea level was attained (about 6.5 ka), has been deduced by many researchers from lake levels and salinities, from vegetation signatures, and from aeolian dust deposit records. In general, from about 8 to 5 ka, the climate was wetter than present (Bowler 1981; Dodson 1986; Chivas et al. 1985, 1993; Magee et al. 1995; Anker et al. 2001). From about 5 to 2 ka, conditions were drier than present with significant changes in Tasmanian pollen spectra (Anker et al. 2001) and changes in inland lakes and sediments (e.g. Stanley and De Deckker 2002). Chivas et al. (1985) and Bowler (1981) showed that from about 2000 BP to 300 or 400 BE Lake Keilambete in Victoria returned to perennial lake conditions. However, this return of high lake levels was not shown by Dodson (1986) for Breadalbane or Coventry and Walker (1977) at nearby Lake George, NSW. It is unclear exactly what natural vegetation changes have occurred on coastal floodplains over the past 6.5 ka in response to changes that occurred in climate and associated evapotranspiration-driven drainage; it is likely that they have been profound. The shift to drier conditions over much of the last few thousand years, compared with conditions initially experienced by landscapes accumulating sulfidic sediments, favours natural landscape drainage. This is particularly the case when base level is also lowered.
It seems to us that natural processes of landscape/hydrology evolution and pedogenesis can account for the existence and degree of much of east-coast Australia's backswamp ASS acidification. Artificial drainage systems may not have caused the acidity formation but they do provide the conduit for its enhanced export. Therefore, if any such landscapes have not been drained, any proposal to initiate their drainage should be avoided.
For those backswamps already drained, the management of the acidity export is essential and must focus on mechanisms to maximise acidity retention in the landscape, and treatment of any acidity being exported in drain systems.
The past and current management of ASS at McLeod's Creek has involved the interception and diversion of hill-slope runoff to bypass the floodplain in major drains. Within the floodplain, management involves the containment of the acidity and other oxidation products through laser-levelling to minimise rain-fed infiltration, and minimisation of lateral drains that arc the major source of acidity export. This approach also aims to minimise acidic export by reducing near-surface soil moisture, and by doing so, increasing the storage for local rainfall whilst maintaining the watertable above the potential ASS layer in order to minimise further sulfidic oxidation (Atkinson and Tulau 2000; Tulau 2002). This appears to be the only viable option for the study area, as the complete neutralisation of the existing acidity store (~50 t [H.sub.2]S[O.sub.4]/ha.) with lime is too expensive for broadacre agriculture (White et al. 1997), and reflooding with brackish river water in an attempt to return the landscape to its original state is clearly socially and economically incompatible with the current landuse.
There are several aspects to the containment strategy that will be discussed in light of the presented results, as well as the methods currently employed at the study site. The drainage system design (i.e. field drain frequency) has been shown to be of considerable importance (Yang et al. 2000). The results here, and other work (Smith et al. 2003), show that there is an extensive pool of stored acidity within these canefields, and that the current drainage network provides the conduit for exporting such oxidation products from the system. The results also suggest that a large proportion of the oxidation products being exported to field drains are doing so close to the field drain edge. However, even though it appears that drain edges are an important factor in the discharge of oxidation products from ASS, complete removal of the majority of field drains is not the panacea for estuarine degradation.
The major and minor field drains are required for the discharge of mole drains (Rittie 2000) and the storage of runoff. The drainage system managed as it is in the 100-ha farm on which this study was completed can store about 25 mm of rainfall before water enters the lowermost cane-growing surfaces. An excessive reduction in drain storage capacity would result in increased infiltration previously removed as runoff, and ponding/waterlogging in lowest landscape positions during periods of rain (White et al. 1997). In the absence of laser-levelling, additional infiltration would mean an increase in the volume of water passing through the acidified zone, allowing for extensive ion exchange and acidification of through-flow waters. Furthermore, the fields surrounding the study area are amongst the most productive in NSW (White et al. 1993), despite storing large amounts of acidity. Therefore, a balance needs to be achieved between minimising the number of field drains to contain the acidity within the cane fields, and still fulfilling the above-mentioned hydrological functions to enable the maximum growth of the sugarcane. Infilling the equivalent of about half the total length of field drains at McLeod's Creek is being carried out in an attempt to decrease the acid export from the site (Smith et al. 2003).
As part of the drainage system design, the depth of the field drains should also not extend into the unoxidised material (White et al. 1996). The results here show that the sulfidic sediment is approximately 0.4 m below the bottom of the field drain. Therefore, care should be taken if cleaning the drain not to excavate any lower than the current depth, as this will provide a new surface of oxidation when the drain water levels are reduced in low rainfall periods, and a rapid transfer path for oxidation products out of the system when local rainfall does increase (White et al. 1996, 1997). The application of lime to drains prior to their cleaning is also undertaken.
The final component of the acidity containment strategy is the active control of the water level in the drain system, and the watertable under the growing crop. A low watertable beneath the crop but above the sulfidic layer has the ability to store water from rainfall, removing it from the surface water system and subsequently removing it through evapotranspiration (Wilson et al. 1999). Watertable management has been performed at McLeod's Creek for that past 17 years, where the water level in drains is maintained at -0.6 m AHD at the outlet drain with an electric pump (Smith et al. 2003). This, along with the partial opening of floodgates to let brackish water (about 1/2 seawater salt concentration) enter the drains during drier periods, has ensured that the watertable has remained above the sulfidic layer whilst ensuring maximum storage capacity for local rainfall (Smith et al. 2003), as can be seen in the reported data (Fig. 6).
This strongly brackish water is allowed into the 100-ha farm drainage system during very dry periods to ensure that monosulfidic drain-bottom sediments remain saturated. These sediments store acidity, dissolved metals, and other pollutants so that they can later be safely removed and lime-treated before spreading on adjacent cane paddocks. The input of this brackish water is carefully managed so that it does not enter mole drains where it would greatly impact the salt-sensitive cane crop. That such salty water input into the drains is possible clearly shows that the lateral hydraulic conductivity is very small in most of the sulfuric and all of the sulfidic layers of these ASS. This management technique has also been used in other Tweed cane-growing areas but its general application depends on a careful hydraulic assessment of each area.
White et al. (1993) showed from movement of Tweed brackish water up Mcleod's Creek during dry conditions that there was some connection between crop evapotranspiration and the drainwater. Nevertheless, as Wilson et al. (1999) showed, this linkage is not strong, so that the crop depends on rain storage from vertical infiltration at the immediate site. Groundwater from adjacent hills is excluded from the crop by the depth (~10 m) of sulfidic gel beneath the crop, acting as an aquaclude.
The large variability in rainfall at the study site means that the acidity containment strategy described here may occasionally fall short of preventing large discharges of acidity from occurring. There is therefore the potential for neutralising the discharge waters at such times with lime (agricultural, hydrated or dolomitic) via a Calibrated Reagent Applicating Blender (CRAB) (Desmier et al. 2003) after any significant discharge events. Also, as the results show that a large proportion of the acidity and oxidation products is coming from only a small portion of the canefield, the strategic application of lime to canefield headlands, field drain banks, and their mole drain exit points would be effective in targeting the highest areas of acidity release, compared with the general application of lime across an entire canefield.
Whilst these suggestions may be applicable to areas with similar depths to sulfidic sediments and climatic conditions, they may not translate to all areas affected by ASS under sugarcane cropping. The management of acidic discharges from broad-scale, low-value landuses other than sugarcane remains problematic. The specific parameters of individual sites need to be taken into consideration before detailed management strategies are implemented.
The authors would like to thank landowner Robert Quirk for his continued support and participation with our research into ASS. The provision of funding by the Department of Land and Water Conservation (ASSPRO Grant and ASS Encouragement Award) is acknowledged. The authors also wish to acknowledge Jason Reynolds for many helpful discussions.
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A. S. Kinsela (A,B) and M. D. Melville (A)
(A) School of Biological, Earth and Environmental Sciences, University of New South Wales,
Sydney, NSW 2052, Australia.
(B) Corresponding author; email: firstname.lastname@example.org
Manuscript received 16 May 2003, accepted 31 March 2004
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|Author:||Kinsela, A.S.; Melville, M.D.|
|Publication:||Australian Journal of Soil Research|
|Date:||Sep 1, 2004|
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