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Chemical changes during oxidation of iron monosulfide-rich sediments.


Iron monosulfides (FeS) are an important component in the cycling of iron and sulfur within acid sulfate soil (ASS) environments and occur either as coatings on pyrite framboids (Bush and Sullivan 1999) or as bottom sediments in drains and streams as the commonly known 'monosulfidic black oozes (MBOs)' (Sullivan et al. 2002). Iron monosulfides vary in form and include amorphous FeS (Fe[S.sub.amorph]), mackinawite (tetragonal Fe[S.sub.1-x]), and greigite ([Fe.sub.3] [S.sub.4]) (Morse et al. 1987). Iron monosulfides are the initial iron sulfide minerals formed under reducing conditions in estuarine sediments, and are often necessary precursors for the rapid formation of pyrite (Berner 1970).

In ASS environments, reducing conditions may develop in drains during periods of nil or low flows. These conditions, plus sufficient dissolved sulfate and organic matter, favour the formation and accumulation of FeS in the drain sediment. in most anoxic sediments, sulfides are considered the predominant solid phase controlling the concentrations of the metals Cd, Cu, Fe, Mn, Ni, Pb, and Zn (Simpson et al. 1998). Adsorption of chemicals onto the sediments represents an important sink fur contaminants (including acid, iron, and aluminium), carbon and nutrients (Jickells and Rae 1997), as well as heavy metals. Iron monosulfide formation is therefore important in the improvement of acid drainage from ASS-affected areas.

The critical issue with sulfidic anoxic sediments is the potential for oxidative release of metals and other contaminants that might occur by physical disturbances such as dredging or drain cleaning, strong currents, or the introduction of oxic waters at depth by burrowing organisms (Simpson et al. 1998). Upon oxidation, these sulfides may oxidise, producing oxidised sulfur species (i.e. S[O.sup.2-.sub.4], [S.sup.0]), and releasing associated metals into the water column (Simpson et al. 1998). Oxidation can also occur during droughts when drains dry out, exposing the bottom sediments to air. Subsequent rainfall can then mobilise metals. It is likely, especially in estuarine sediments, that the oxidation of sulfide phases represents the major source of metals (Simpson et al. 1998). Iron monosulfides have been identified as a potential environmental hazard in ASS landscapes (Robertson et al. 1998; Sullivan and Bush 2000), including that the oxidation of monosulfides causes de-oxygenation and acidification of drainage waters (Sullivan et al. 2002).

Oxidation of iron monosulfides releases S[O,sup.2-.sub.4] into solution (Bloomfield 1972) and results in the formation of ferrous iron intermediate and acid production (Aller 1980):

(1) 4FeS + 8[O.sub.2] [left arrow] + 4[Fe.sup.2+] + 4S[O.sup.2-.sub.4]

(2) 4[Fe.sup.2+] + [O.sub.2] + 10[H.sub.2]O [left arrow] 8H+ + 4Fe[(OH).sub.3]

The presence of iron monosulfides in ASS environments has important implications for land and drain management (Bush and Sullivan 1998). Avoidance of iron sulfide oxidation is the most suitable management technique in ASS environments. However, this strategy is not always realistic given agricultural and other commercial pressures, and it will not eliminate the natural disturbances to the iron and sulfur cycles. Consequently, active intervention strategies are required to minimise the impact of oxidation products on the environment. ASS materials, including iron monosulfides, that are subject to oxidation are often treated with neutralising agents to minimise the export of acidity. An incubation experiment was conducted to determine chemical changes to iron monosulfide-rich drain sediments from the McLeods Creek floodplain during oxidation. Agricultural lime and mill mud/ash (an alkaline sugar mill byproduct) were applied to the sediments to determine their effectiveness as neutralising agents.



The McLeods Creek floodplain (about 450 ha) is located in the backswamp area of the Tweed River estuarine embayment. The site is used for sugarcane farming and has been extensively drained with > 100 km of drainage channels. Black sediments (MBOs) have been identified in the bottom of all drains across the floodplain. A bulk sample was extracted from a main drain where the MBOs are approximately 0.3 m deep. The sample was stored in a cooler and transported back to Sydney; pH and redox measurements, as well as red iron stains, revealed that the top sediments had oxidised, but this only affected the top few millimetres of sediment, which was subsequently removed. The remaining sediments were black.

The wet sediments were analysed in triplicate for acid-volatile sulfur (AVS) (Bush and Sullivan 1998) and chromium reducible sulfur (CRS) (Sullivan et al. 2000). This CRS method is actually a measure of total reduced inorganic sulfur, including pyrite, FeS, and elemental S. A portion of the unoxidised sediment was oven-dried at 105[degrees]C for 24 h to correct for dry weight. The AVS method of Bush and Sullivan (1998) uses 12 M HCI and extracts monosulfides including amorphous FeS, mackinawite, and greigite. AVS was also determined using 1 M and 6 M HCl. The metals that dissolve during this extraction are known as simultaneously extracted metals (SEM) (Simpson et al. 1998). Because sulfides are important in controlling metal concentrations, the analysis of SEM provides information on the availability of metals in the sediment. In the literature, 1 M HCl has been used for AVS and SEM determinations, so by using 1 M HCl we were able to compare our results with those of other authors, e.g. Allen et al. (1993). SEM was determined by filtering the sediment slurry after the AVS digestion and analysing the extract by ICP-OES for Al, Cd, Co, Cr, Cu, Fe, Mn, Ni, Pb, and Zn. SEM is the sum of the molar concentrations of Cd, Cu, Ni, Pb, and Zn (Allen et al. 1993). A portion of the bulk sediment was oven-dried at 85[degrees]C, crushed, and analysed for total C and S by LECO. Organic matter was determined by loss on ignition (LOI) at 550[degrees]C. The elemental composition of the sediments was measured by ICP-OES following an aqua regia (3 : 1 HCl : HN[O.sub.3]) microwave digestion.

Oxidative incubation experiment

About 1 kg of wet, unoxidised material was placed into each of 9 polyethelene bags for incubation. To 3 of the bags, 100 g of agricultural lime was thoroughly mixed through the wet sediment. To another 3 bags, 100 g of mill mud/ash from the Condong Sugar Mill was thoroughly mixed through the wet sediment. The remaining 3 bags remained untreated. The bags were pressed to form a 1-cm-thick slab, and stored with the ends open in a constant temperature room at 25[degrees]C. This procedure was adopted to approximate slow oxidation similar to oxidation in the field, but without leaching losses of oxidation products (van Breemen 1973). The lime and mill mud/ash used in the experiment were analysed by XRF to determine their composition.

Chemical analysis

Chemical analysis was carried out on the incubation samples over a period of 61 days. On each sampling day, pH and redox was determined by inserting probes connected to a TPS MC-81 at 3 places. Since drying the samples causes chemical changes, all analysis was performed on wet samples and later corrected for dry weight. Sediment : water and sediment : KCI extracts were prepared using approximately 5 g of oven-dried equivalent material and 25 mL of deionised water or 1 M KCl. The extracts were shaken for half an hour and centrifuged at 3000 r.p.m. for 10 min.

The supernatants from the water and KCI extracts were measured for pH and electrical conductivity (water extracts only). Acidity was determined by titration with NaOH to pH 8.3 (APHA 1998). The remaining sample was acidified and analysed by ICP-OES for Al, Ca, Fe, K, Mg, Mn, Na, S, and Si. Sulfate was calculated from the S concentrations. Soluble iron (ferrous and total) was determined on 1 : 1 sediment : water extracts using the colourimetric method of Loeppert and Inskeep (1996).

Many researchers thoroughly mix the materials at each sampling time during incubation experiments (e.g. Crockford and Willett 1995). However, it was thought that this would introduce additional oxygen into the materials. Also, this experiment aimed to mimic oxidation conditions in the field where the monosulfidic drain sediments arc either stockpiled or spread on the paddocks, and then left undisturbed. Each treatment was performed in triplicate and standard errors calculated to account for any unconformities in oxidation.


Bulk sediment characteristics

The AVS concentration of the original sediment was 0.745% SAV or 232.3 [micro]mol S/g, and the CRS concentration was 0.98% [S.sub.CRS]. Therefore the concentration of iron monosulfides in the sediments was approximately 2%, and (assuming elemental S to be insignificant) the pyrite content was 0.44%. The total S concentration was 1.06%, indicating that almost all sulfur exists as reduced inorganic sulfides, mostly in the monosulfide form.

The total C concentration was 4.05% and LOI was approximately 15%. Previous analysis of these sediments has shown that almost all carbon exists as organic carbon with only trace amounts of inorganic carbon (J. Smith, unpublished data). This indicates that the LOI procedure also measures the loss of volatile species such as hydrated clays, and overestimates the amount of organic matter.

The concentration of SEM ([summation over]Cd, Cu, Ni, Pb, Zn) was 2.08 [micro]mol/g with Zn contributing 68%. If the concentration of AVS exceeds the sum of the heavy metal concentrations (SEM), then this excess sulfide is able to bind heavy metals in insoluble and non-bioavailable forms, and therefore the metals will not cause toxicity (ANZECC 2000). The SEM/AVS ratio is very small (0.009), indicating a large excess of monosulfide minerals and that an input of additional metals in the system could easily be locked up in the sediments. Because the metals are bound to the sulfides, their bioavailability is limited and therefore they are not a threat in their natural state. However, it is important to note that upon exposure to air, the sulfide minerals rapidly oxidise and the bound metals could potentially be re-released.

The 3 acid strengths used to extract AVS all gave similar sulfide concentrations with a relative standard deviation (RSD) of only 6.67% as shown in Table 1. However, the SEM results were significantly different (RSD = 37.52%), with 12 M HC1 yielding a greater amount of metals. Similar differences were found in the SEM/AVS ratio (RSD = 44.18%). Although the different acid concentrations affect the amount of metals extracted from the sediment, the sulfide content is unaffected. Therefore, if 12 M HCI is used, the only effect will be an overestimation of the metal bioavailability and the SEM/AVS ratio becomes closer to one. This ratio is often used in sediment quality guidelines (ANZECC 2000), and in terms of management, the overestimation of metal bioavailability will simply mean a more precautionary approach will be taken. The small difference in the AVS concentration between the 3 acid strengths indicates a lack of crystalline monosulfides such as greigite in the sediment.

The aqua regia digest results are presented in Table 2. Also included are the ANZECC sediment quality guidelines for certain metals (ANZECC 2000). The 2 values are the low trigger and high trigger values. The sediments contain metal concentrations below the low trigger values and are therefore considered low risk.

The sediments contain large amounts of aluminium, iron, and sulfur. Unfortunately, there is no standard guideline in Australia for these elements to determine if they are in toxic quantities. There are also many other metals which may potentially be hazardous, including cobalt and manganese.

Chemical changes during oxidation

The unoxidised sediment was of pH 6.75, which was similar to values measured in the field. The initial pH of the 3 treatments was 6.58, 6.58, and 6.51 (untreated, lime, and mill mud/ash, respectively). Over the 61-day oxidation period, the pH in the untreated samples decreased to 3.51. The pH of the mill mud/ash treated sample decreased to 5.40, and the pH of the lime-treated sample increased to 6.82, as shown in Fig. 1. The EC values during the experiment remained fairly constant, at 4-8 dS/m for all 3 treatments.


Changes in redox during the oxidation of the sediments is shown in Fig. 2. The Eh of the untreated sediments increased gradually from--265 to 355 mV. The Eh of the limetreated sediments also increased from-278 mV but remained negative with a final value of-36 mV. The Eh of mill mud/ash treated sediments actually decreased from -267 to -458 mV before gradually increasing to -199 mV at the end of the study period. The Eh in the mill mud/treated samples began to increase after 12 days, indicating the onset of oxidation. It appears that mill mud/ash and, to a lesser extent, lime inhibit the oxidation of the sediments, thereby reducing the effects of oxidation such as acid generation.


Soluble and exchangeable acidity in the untreated samples was considerably greater than the lime and mill mud/ash treated samples. Figure 3 shows that soluble plus exchangeable acidity in the untreated samples increased to a maximum of 5.4 g/kg before decreasing to 4.3 g/kg of CaC[O.sub.3] equivalent at the end of the study period. Acidity in the lime-treated samples was initially 0.5 g/kg but remained <0.4 g/kg of CaC[O.sub.3] equivalent for the entire study period. In the mill mud/ash treated samples, the acidity decreased from 0.9 to <0.2 g/kg before gradually increasing to 1.7 g/kg of CaC[O.sub.3] equivalent after 61 days.


Changes in soluble ferrous iron concentrations followed similar trends to soluble acidity. Concentrations in the lime-treated samples remained <1 [mmol.sub.c]/kg for the entire period. In the mill mud/ash treated samples, the concentration increased from 1.7 to 2.5 [mmol.sub.c]/kg, then decreased to 0.4 [mmol.sub.c]/kg before increasing to 5.7 [mmol.sub.c]/kg, as shown in Fig. 4. The ferrous iron concentration in the untreated samples remained <5 mmolc/kg for the first 14 days, then increased rapidly to 32.3 mmolc/kg after 33 days and decreased to 4.6 [mmol.sub.c]/kg after 61 days. Total soluble iron concentrations were similar to soluble ferrous iron concentrations, indicating that soluble iron was in the reduced form. Any ferric iron produced subsequently formed insoluble iron oxides/oxyhydroxides and this was evident by the increasing red colour of the sediments during oxidation.


Soluble sulfate in the untreated and lime-treated samples increased gradually from approximately 30 to 220 [mmol.sub.c]/kg (Fig. 5). The decrease in CRS (total reduced sulfide) or AVS after oxidation was not measured, but since almost all sulfur exists as sulfides, mainly in the form of monosulfides, the source of this sulfate is from sulfide oxidation. The fact that sulfate concentrations also increase in the lime-treated samples indicates that it does not actually prevent sediment oxidation. However, the high pH and low acidity concentrations in the lime-treated samples show that the lime neutralises any acid produced during oxidation.


In the mill mud/ash treated samples, the soluble sulfate concentration actually decreased to almost zero, then after 33 days began increasing to 156 [mmol.sub.c]/kg. The initial decrease in sulfate indicates sulfate reduction, which is coupled to very low Eh values. There is also a concurrent decrease in acidity and ferrous iron, which is a result of alkalinity production during sulfate reduction and possible FeS precipitation. It appears that mill mud/ash actually inhibits initial sediment oxidation, or at least binds all oxidation products, but its capacity is limited, and after about 30 days, sediment oxidation begins or its binding capacity is exhausted.

There are strong positive relationships between soluble acidity and soluble sulfate ([r.sup.2] = 0.83) and soluble acidity and soluble ferrous iron ([r.sup.2] = 0.80) in the untreated samples indicating that the majority of acidity is from iron sulfide oxidation. The precipitation of insoluble iron oxides/oxyhydroxides controls the [Fe.sup.2+] concentrations and subsequently reduces the acidity concentration.

The release of aluminium, manganese, and silica during sediment oxidation is shown in Figs 6, 7, and 8. Oxidation of the untreated samples released aluminium, manganese, and silica, indicating the breakdown of aluminosilicate minerals (clays) and the release of heavy metals by acid attack. Mill mud/ash appears to have limited neutralisation capacity, since after 30 days, soluble manganese and silica increased. However, soluble aluminium decreased after 30 days, possibly due to complexation or binding to organic matter. Lime appears to be an effective neutralising agent, since aluminium, manganese, and silica concentrations remained close to background levels for the entire study period.


The main elemental components of the lime and mill mud/ash are shown in Table 3. The lime consisted mostly of calcium and silica, and the mill mud/ash consisted mostly if silica, aluminium, and iron. It is assumed that there is a large proportion of aluminosilicate minerals in the mill mud/ash, and these, as with the untreated samples, are subject to attack by acid during sediment oxidation. This explains the limited effectiveness of mill mud/ash as a neutralising agent.


Leonard et al. (1993) describe the AVS concentrations commonly reported in the literature. For example, in unpolluted freshwater sediments, AVS has been reported in the range of about 4-13 [micro]mol S/g dry weight, while AVS concentrations in coastal marine sediment have been reported in the range of 20-90 [micro]mol S/g dry weight. In different depositional environments, there is a tendency for different components to be limiting. Sulfate limitations are the major cause of lower concentrations of AVS in freshwater v. marine sediments (Leonard et al. 1993).

In ASS environments, AVS concentrations can be quite variable. For example, Claff et al. (2002) report AVS concentrations ranging from 63 to 1005 [micro]mol/g in the Tuckean Swamp in the Richmond River catchment, and 11 to 375 [micro]mol/g in the Clarence River catchment. The AVS concentrations measured in the drains of McLeods Creek are therefore much greater than typical freshwater and marine sediments, but are much smaller than in some other ASS environments in Australia.

These large concentrations of iron monosulfides in ASS environments play an important role in the binding of heavy metals, van den Berg et al. (1998) found that AVS concentrations as low as 0.02% were large enough to bind significant amounts of heavy metals.

The aqua regia digestion of sediments does not fully break down all minerals in the sediments, but instead provides a worst-case scenario if the sediments were subject to severe acid attack. The results have shown that heavy metal concentrations do not exceed the ANZECC sediment quality guidelines. However, the guidelines do not provide information for many of the metals released from iron monosulfides, especially iron and aluminium. Elevated iron and aluminium levels in discharge waters from ASS have been reported as a major factor affecting water quality and contributing to fish kills (Easton 1989). Whilst the research in this area has focussed on iron and aluminium being released from actual ASS, the oxidation of iron monosulfides andoubtedly contributes to these elevated concentrations.

The results indicate that the chemical changes observed during sediment oxidation, including increased acidity, [Fe.sup.2+], and S[O.sup.2-.sub.4] and decreased pH, are a result of iron sulfide oxidation. In these sediments, the pyrite concentration is small (<20% of CRS) compared with the monosulfide content. Additionally, iron monosulfides are far more reactive than pyrite, and the oxidation of iron monosulfides proceeds much more rapidly than that of pyrite and readily takes place abiogenically at high pH (Aller 1980). Therefore, while changes in iron monosulfide and pyrite concentrations were not measured during sediment incubation, it is concluded that most of the chemical changes observed are a result of iron monosulfide oxidation, especially in the initial stages of oxidation. However, it is likely that the oxidation of pyrite does partially contribute to the chemical changes observed. This conclusion was also reached by Saulnier and Mucci (2000), who found the release of Fe and Mn to the dissolved phase during sediment resuspension experiments likely originates from the oxidation of AVS (i.e. monosulfides).

During oxidation of iron monosulfides, dissolved oxygen is consumed, producing ferrous iron and zero-valent sulfur (Bush et al. 2002). Acid production occurs after the dissolved oxygen concentration rises sufficiently to allow the acid-producing oxidation of ferrous iron ([Fe.sup.2+]) to ferric iron ([Fe.sup.3+]) and sulfur to sulfate. The oxidation of [Fe.sup.2+] to [Fe.sup.3+] occurs slowly, and at pH 7, [Fe.sup.2+] may be expected to persist for some time (Simpson et al. 1998). Peterscn et al. (1997) found that certain particulate-bound trace elements can be released during the re-oxidation of anoxic sediments and that this release is strongly affected by microbial processes. The oxidation rate of sulfide is increased by the presence of metal ions (especially [Fe.sup.3+]). Thus, it is possible that initial oxidation was slow, but given sufficient bacterial input and production of [Fe.sup.3+], the rate of oxidation rapidly increased. This explains the delay in acidification observed during sediment oxidation.

Iron monosulfides can produce acid by oxidation as well as acidification. FeS is acid-soluble and will easily dissolve according to the following reaction:

(3) FeS + [H.sup.+] [left arrow] [Fe.sup.2+] + HS-

This means that in an acid discharge event from the surrounding ASS, monosulfides in the drains could easily dissolve and the release of [Fe.sup.2+], providing the opportunity for even more acid production. The production of extra protons continues to dissolve FeS (Schippers and Jorgensen 2002).

Although the purpose of the addition of neutralising agents is primarily to neutralise the acidity formed by oxidation of sulfides (both by [O.sub.2] and [Fe.sup.3+]), the accompanying increase in pH also inhibits microbially mediated oxidation (Arkesteyn 1980) and lowers the solubility of [Fe.sup.3+] (Ward et al. 2002). The addition of carbonate materials aids in the formation of iron oxide coatings on sediment surfaces, which inhibits oxygen diffusion (Ward et al. 2002). The application of mill mud/ash appears to bind acidity and metals due to the large amount of organic matter, but its neutralisation potential is limited. In the absence of neutralising materials, low pH conditions persist due to strong buffering exerted by the hydrolysis of aluminosilicates and iron in the system (Gurung et al. 2000).

Disturbance of sediments at McLeods Creek usually occurs during drain cleaning operations by land managers, or if the drains dry out. At other ASS sites, e.g. in the Richmond River catchment, monosulfides are much more mobile and may be disturbed during large rainfall events. Drain cleaning is an inevitable practice in sugarcane farming. However, it is important to recognise the hazards involved and therefore look at ways of minimising them. Typically the sediment is removed with a scraper bucket and spread across adjacent paddocks where it is mixed with lime. Iron monosulfide-rich drain sediments that have not been treated with lime have been reported to affect cane production by up to 30% (R. Hawken, pers. comm.). The only certain method of preventing oxidation and acidification is to maintain sediments under anaerobic conditions (Bloomfield 1972).

It is clear that the iron monosulfide-rich sediments have the potential to cause environmental damage and present a threat to water quality in ASS landscapes. In their summary of MBO distribution and behaviour, Sullivan et al. (2002) explain that the environmental concerns related to MBOs are partly based on the high reactivity of iron monosulfides and that appreciable concentrations of iron monosulfides can occur in these oozes. Acid production from MBOs has been considered to be appreciable, and accordingly, it is now a best practice requirement for NSW canefarmers to add lime to these materials when they are excavated.

The installation of a new saltshaker floodgate into the western subcatchment of McLeods Creek allows an input of saltwater into the drainage system to prevent the drains drying out. The floodgate has the benefits of adding seawater to neutralise acidity and maintain sufficient water levels to inhibit oxidation. However, it also supplies additional sulfate, which is the limiting factor in monosulfide formation in this system and thereby improves water quality by locking up iron and other contaminants in ASS discharge waters.

It appears that lime is a more effective neutralising agent than mill mud/ash. The 2 neutralising agents were applied in quantities thought to be in excess of the acidity expected, in order to gain preliminary information about their effectiveness and impact. A study completed by Toppler (2003) involved the application of various rates of lime and mill mud/ash to actual ASS materials. However, further trials involving various application rates of lime and mill mud/ash need to be conducted, and the sulfide species changes and the costs and availabilities need to be evaluated before any conclusions can be made regarding the most effective treatment. However, it is clear that some kind of neutralising agent is essential to ensure that cane production and water quality are not affected by sulfide oxidation products.


The role of iron monosulfides as a sink for contaminants is important in improving water quality from ASS discharges. However, this study shows that the oxidation of iron monosulfide-rich sediments causes important chemical changes, such as the release of metals and production of acidity. If left untreated, large amounts of iron and aluminium are produced, and these may contribute to acidic discharges and fish kills. Due to the large concentration and highly reactive nature of iron monosulfides in these types of sediments, it is concluded that most of the chemical changes observed are the result of iron monosulfide oxidation. However, the contribution of pyrite oxidation to these chemical changes cannot be excluded and measurement of all sulfide species should be undertaken during such experimentation. Treatment of oxidising sediments with lime appears to be effective in controlling the impacts of oxidation. Mill mud/ash also has a limited use as a neutralising agent, although further studies are required to fully assess its potential. These findings highlight the need for best management of drains in ASS areas.
Table 1. Comparison of AVS and SEM results determined with different
HCl concentrations

([micro]mol/g) ([micro]mol/g)

1 M 233.98 1.50 0.0064
6 M 246.93 1.77 0.0072
12 m 216.08 2.96 0.0137
Mean 232.33 2.08 0.009
RSD (%) 6.67 37.52 44.18

Table 2. Sediment composition determined by aqua regia digestion

Element mg/kg ANZECC

Al 60367.98 --
As < 0.03 --
Au < 0.02 --
Ba 85.35 --
Ca 3283.73 --
Cd < 0.001 1.5-10
Co 33.09 --
Cr 59.81 80-370
Cu 23.22 65-270
Fe 82855.12 --
K 4440.09 --
Li 52.54 --
Mg 5660.74 --
Mn 9.18 --
Mo 0.27 --
Na 7439.15 --
P 862.08 --
Pb 11.32 50-220
Si 600.77 --
S 12862.71 --
Sn 9.58 --
Sr 70.15 --
Ti 1108.19 --
W 2.57 --
Zn 143.51 200-410

Table 3. XRF analysis results for lime and mill mud/ash

Compound Lime mud/ash

[Na.sub.2]O 0.45 1.08
MgO 0.74 1.52
[Al.sub.2][O.sub.3] -0.37 9.5
Si[O.sub.2] 2.5 41.91
[P.sub.2][O.sub.5] 0.049 2.28
S[O.sub.3] 0.323 0.33
[K.sub.2]O 0.182 1.38
CaO 54.76 3.09
Ti[O.sub.2] 0.033 1.09
[V.sub.2][O.sub.5] 0.011 0.02
[Cr.sub.2][O.sub.3] 0.008 0.03
MnO 0.027 0.13
[Fe.sub.2][O.sub.3] 0.92 4.36
NiO -0.009 0
CuO 0.006 0.01

ZnO 0.003 0.02
[As.sub.2][O.sub.3] -0.007 -0.01
[Rb.sub.2]O -0.011 0
SrO 0.007 0.01
[Y.sub.2][O.sub.3] 0.002 0
Zr[O.sub.2] 0.007 0.02
BaO -0.056 0
Ce[O.sub.2] 0.079 0.08
PbO 0.019 0.02
Th[O.sub.2] 0.132 0.05
[U.sub.3][O.sub.8] 0.023 0.01
Total XRF 59.83 66.92
LOI 39.15 33.54
Total 98.98 100.46


This research project was undertaken as part of Project ASS 00.35 funded by the NSW Department of Agriculture Acid Sulfate Soils Program (ASSPRO), and assisted by a Commonwealth of Australia--Australian Postgraduate Award.


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School of Biological, Earth and Environmental Sciences, University of New South Wales,

Sydney, 2052, Australia; email:

Manuscript received 16 May 2003, accepted 3 May 2004
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Author:Smith, J.
Publication:Australian Journal of Soil Research
Geographic Code:8AUST
Date:Sep 1, 2004
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