Fractionation and extractability of sulfur, iron and trace elements in sulfidic sediments

Chemosphere 64 (2006) 1421–1428 www.elsevier.com/locate/chemosphere

Fractionation and extractability of sulfur, iron and trace elements in sulfidic sediments Edward D. Burton *, Richard T. Bush, Leigh A. Sullivan Centre for Acid Sulfate Soil Research, School of Environmental Science and Management, Southern Cross University, P.O. Box 157, Lismore, NSW 2480, Australia Received 30 September 2005; received in revised form 24 November 2005; accepted 1 December 2005 Available online 24 January 2006

Abstract This study describes iron and sulfur fractionation, and the related extractability of selected trace elements (As, Cd, Cr, Cu, Ni, Pb and Zn) in estuarine sediments. The sediments were sulfidic, with moderately high concentrations of pore-water sulfide (200–600 lmol l1) and acid-volatile sulfide (AVS; 9.9–129 lmol g1). Pyrite-S concentrations increased with depth, with 63–251 lmol g1 at site W1 and 312–669 lmol g1 at site W2. The degree of sulfidisation was generally high (>80%), indicating that Fe may be limiting pyrite accumulation. The ratios of AVS to pyrite-S increased with sediment depth, as expected for the pyritisation of solid-phase AVS. Cadmium, Pb and Zn extractability in 1 M HCl indicated that these elements are not significantly sequestered during pyritisation, whereas sequestration may be important for As, Cu and possibly Ni. Extractability trends for Cr suggest that diagenesis in sulfidic sediments may enhance Cr reactivity. Overall, replacement of AVS by pyrite during diagenesis may enhance the reactivity of Cd, Cr, Pb and Zn, whereas As, Cu and possibly Ni may be rendered less reactive. Ó 2005 Elsevier Ltd. All rights reserved. Keywords: Pyrite; AVS; Heavy metals; Diagenesis

1. Introduction Pyrite and other reactive iron-sulfide minerals are important to sedimentary trace element behaviour (Morse and Luther, 1999). These minerals provide sinks for potentially toxic trace elements that are introduced to estuarine systems from anthropogenic and natural sources (Chapman et al., 1998; Simpson et al., 2002). The bioavailability and potential mobility of trace elements in sulfidic sediments is therefore largely determined by the geochemical cycling of key Fe and S fractions (Cooper and Morse, 1998a; Teasdale et al., 2003). Acid-volatile sulfide (AVS) is a major sorbent for several trace elements in anoxic sediments (Morse and Arakaki, 1993). Several studies show that a molar excess of AVS

*

Corresponding author. Tel.: +61 2 6620 3638. E-mail address: [email protected] (E.D. Burton).

0045-6535/$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2005.12.003

over reactive trace metals (defined as extractable in 1 M or 6 M HCl) indicates low metal availability (Di Toro et al., 1992; Chapman et al., 1998). In this situation, a relative abundance of reactive sulfides effectively sequesters reactive trace elements, thereby reducing bioavailability. As such, comparison of trace element extractability with AVS levels is an important part of sediment quality assessment (Ankley et al., 1994; ANZECC/ARMCANZ, 2000; Burton et al., 2005). Acid-volatile sulfides comprise pore-water sulfides and a suite of meta-stable iron-sulfide minerals, such as amorphous FeS, mackinawite and greigite (Morse and Rickard, 2004). Over time, sedimentary AVS tend to be replaced by thermodynamically favoured pyrite (Morse et al., 1987). The replacement of solid-phase AVS (e.g. amorphous FeS) by pyrite does not involve a literal ‘‘conversion’’ of FeS to pyrite (Morse and Rickard, 2004). Rather, FeS dissolves and pyrite forms, resulting in increasing pyrite-S to AVS ratios with time (Luther, 1991; Morse and Rickard,

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2004). As FeS dissolves, any associated trace elements will also be liberated. Acid-volatile sulfide therefore exerts only a temporary control on the availability and mobility of trace elements. The ability of sulfidic sediments to provide a long-term trace element sink is therefore dependent on trace element sequestration during pyritisation. Acid-volatile sulfide is operationally defined as reduced inorganic S that reacts with HCl to form H2S (Morse and Cornwell, 1987; Di Toro et al., 1992; Allen et al., 1993; Hsieh et al., 2002). Reactive Fe and trace elements have also been defined by HCl extraction (Canfield, 1989; Simpson et al., 2002; Teasdale et al., 2003; Burton et al., 2005). Similarly, the pyrite-bound fraction has been defined as trace element sulfide phases that are not HCl-extractable over short time-frames (Huerta-Diaz and Morse, 1990, 1992). As a consequence, the recovery of trace elements by HCl extraction may be expected to decrease as the ratio of AVS to pyrite-S decreases. In this contribution, we describe Fe–S fractionation and the related extractability of selected trace elements (As, Cd, Cr, Cu, Ni, Pb and Zn) in sulfidic, estuarine sediments. In particular, we examine trace element extractability as influenced by Fe–S diagenesis in sediments containing moderately high levels of pore-water sulfide. 2. Methods 2.1. Environmental setting The study sites were located in an inter-tidal, sub-tropical, Avicennia-dominated mangrove forest in south-east Queensland, Australia (153,10.08°E; 27,25.34°S) (Fig. 1). The mangrove forest spans 200 m between a former municipal landfill and Moreton Bay, an ecologically sensitive estuary. Previous work has shown that ingress of landfill leachate and deposition of metalliferous materials have lead to sedimentary trace element enrichment within the mangrove forest (Clark et al., 1997; Clark, 1998). Two study sites (termed W1 and W2) were selected based on previous research (Clark et al., 1997; Burton et al., 2005). Site W1 was within a narrow zone (20 m) containing dead mangrove stumps between the mangrove forest and the landfill face. Site W2 was located in a small drainage depression surrounded by mangrove trees, 20 m from inter-tidal, mud-flats of Moreton Bay. 2.2. Sediment collection and handling Sediment cores (internal diameter of 10 cm) were retrieved using a push-tube coring device. The intact cores were extruded step-wise and sectioned, within 4 h of collection, into 2 cm or 4 cm segments. The depth segments were sectioned rapidly (<2 min) and immediately placed into 50 ml polypropylene vials. The vials were completely filled with sediment (no bubbles or headspace) and sealed with gas-tight screw-caps. Possible oxidation of reduced species was minimised by rapid sediment sectioning, avoidance of

Fig. 1. Locations of sites W1 and W2 described in this study.

unnecessary sediment disturbance, and by transporting the sediment-filled vials on ice under an N2 atmosphere. 2.3. Analyses All laboratory glass- and plastic-ware was cleaned by soaking in 5% (v/v) HNO3 for at least 24 h, followed by repeated rinsing with deionised water. All chemicals were analytical reagent grade, and all reagent solutions were prepared with deionised water (milliQ). Sediment moisture content was determined by weight loss due to drying at 105 °C. Sediment pH and Eh were determined using a Beckman U50 m (within 4 h of sample collection) by direct insertion of calibrated electrodes into the sediment sample (Burton et al., 2005). Total C and S was determined on oven-dry (105 °C, 24 h) sediment samples using a LECO-CNS 2000 induction furnace analyzer. Pore-water was extracted within 24 h of sediment collection by centrifugation (2000 g, 20 min) of sediment-filled 50 ml polypropylene vials. The displaced pore-water was

E.D. Burton et al. / Chemosphere 64 (2006) 1421–1428

passed through a 0.45 lm syringe-driven filter. Pore-water 2 sulfide (which includes H2S, HS, SP and aqueous sulfide complexes; collectively denoted as S(-II)(aq)) was immediately preserved using ZnOAc (APHA, 1998), prior to determination by the methylene blue method (Cline, 1969). Pore-water FeTotal and Fe(II) were determined using the 1,10-phenanthroline P method (APHA, 1998). The analytical precision for S(-II)(aq), FeTotal and Fe(III) was within 5%, as determined by duplicate analysis of 25% of pore-water samples. Pore-water retrieval, filtration and mixing with colorimetric reagents was performed in less than 60 s for each sample to limit potential oxidation of reduced pore-water species due to atmospheric exposure. Solid-phase, reduced inorganic S (RIS) fractions were sequentially extracted in duplicate from sediment samples that had been stored frozen, under N2, for no longer than 4 weeks. Acid-volatile sulfide was extracted using the cold diffusion procedure (Hsieh et al., 2002), with the use of ascorbic acid to prevent interferences from Fe(III). The sediment slurry remaining after AVS extraction was transferred to an acetone-washed 50 ml vial, centrifuged (3000 g, 10 min) and the supernatant discarded. Elemental S ðS08 ðsÞ Þ was extracted by shaking the sediment with 20 ml of acetone for 24 h (Wieder et al., 1985), followed by a further 10 ml acetone rinse. The S08 ðsÞ content of the acetone phase (separated by centrifugation at 3000 g, 10 min) was determined by the Cr-reduction method presented by Sullivan et al. (2000). Pyrite-S, in the AVS- and S08 ðsÞ -extracted sediment, was then determined by Cr-reduction analysis (Sullivan et al., 2000). The quantity of S for each solid-phase RIS fractions was determined by iodometric titration (APHA, 1998). Reactive Fe and trace element fractions were extracted by shaking 3 g (wet weight) of sediment with 40 ml of 1 M HCl for 1 h (Burton et al., 2005). Duplicate extractions were performed on 25% of samples and showed that precision was within 10% for all elements except for Cu (precision within 25%, due to low extractable Cu concentrations). Near-total Fe and trace element content was determined by aqua-regia digestion (1:3 HNO3:HCl, 20 min, 1000 W microwave at 10% power) (Simpson et al., 1998). Triplicate digestion of a certified reference sediment (PACS-2, from the National Research Council of Canada) yielded precision within 5% for all elements, except Cd (10%) and Fe (9%). The recoveries of Cd, Cu, Ni and Zn were within the 95% confidence limits of the certified values. The recoveries of the other elements examined here differed only slightly from the certified values for PACS-2 (76% for Cr, 85% for Pb, 79% for Fe, and 109% for As). As a further quality assurance measure, duplicate digestions were performed on 25% of samples with precision found to be within 5% (except for Cd, which was within 12%). Arsenic, Cd, Cu, Cr, Fe, Ni, Pb and Zn concentrations for 1 M HCl extracts and aqua-regia digests were determined by inductively coupled plasma-atomic emission spectrometry (ICP-AES; Perkin–Elmer DV4300) or inductively coupled plasma-mass spectrometry (ICP-MS;

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Perkin–Elmer Optima 4000/30 DV), calibrated using commercially available standards. For these determinations, all samples were first analysed by ICP-AES; samples containing less than the detection limit for ICP-AES were then analysed by ICP-MS. 3. Results and discussion 3.1. General sediment properties Sediment pH was near-neutral, spanning pH 7.07–7.27 at site W1 and pH 6.53–7.46 (Table 1). Neither site W1 or W2 contained living mangrove roots, which are known to promote oxic conditions within mangrove sediment profiles (Clark et al., 1998). The measured Eh values were indicative of strongly reducing conditions (generally <120 mV). Total C contents were high in near-surface sediment (15% for site W1 and 12% for site W2), decreasing to 5–6% at depths of 14–22 cm below the sediment–water interface (Table 1). 3.2. Sulfur and iron fractionation P Pore-water S(-II)(aq) concentrations were moderately high, ranging between 200 and 600 lmol l1 (Fig. 2). Pore-water Fe(III) was undetectable (<1 lmol l1), whilst Fe(II) was 5.6 lmol l1 near the sediment–water interface, decreasing to undetectable levels at depths below 6 cm (Table 1). These P low Fe(II) levels in the presence of moderately high S(-II)(aq) are consistent with published Fe–S solubility data (Morse P et al., 1987). For example, the observed pore-water S(-II)(aq) concentrations at equilibrium with amorphous FeS, having a solubility product of 102.95 (Ksp = {Fe2+}{HS}/10pH), are expected to buffer

Table 1 Selected properties of the sediment profiles examined in this study Depth interval (cm)

pH

Eh (mV; SHE)

Water content (%)

Total C (%)

Total S (%)

Pore-water Fe(II) (lmol l1)

SITE W1 0–2 7.16 2–4 7.14 4–6 7.07 6–8 7.17 8–10 7.16 10–14 7.14 14–18 7.21 18–22 7.27

164 143 121 144 146 145 150 154

80.2 75.8 72.1 68.7 72.2 75.3 71.6 67.9

15.1 13.5 14.1 9.3 8.0 9.9 6.8 5.9

1.22 1.20 1.06 0.87 0.89 1.38 0.96 0.93

5.6 2.2 <1 <1 <1 <1 <1 <1

SITE W2 0–2 6.53 2–4 6.94 4–6 6.81 6–8 6.99 8–10 7.04 10–14 7.24 14–18 7.24 18–22 7.46

94 194 146 156 163 162 164 167

80.1 64.7 65.4 69.6 67.4 63.9 64.1 68.7

11.8 12.0 10.3 11.4 6.8 5.3 5.1 6.3

1.33 0.99 1.68 2.04 2.76 1.61 1.97 3.08

5.6 2.9 1.4 <1 <1 <1 <1 <1

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E.D. Burton et al. / Chemosphere 64 (2006) 1421–1428 Pore-water sulfide (μmol l -1) 0

200 400 600

Depth (cm)

0

0

200 400 600

5

Pore-water sulfide

10 15

(W2)

(W1)

20

AVS, S80(s) and FeS2-S (μmol g-1) 0

100 200 300

0

30 375

750

Depth (cm)

0 5

AVS S80-S FeS2-S

10 15

(W2)

(W1)

20

AVS / pyrite-S 0

0.8

1.6

Depth (cm)

0

2.4

0

0.04 0.08 0.12

5 10

AVS/pyrite-S

15 20

(W1)

(W2)

greater depth below the sediment–water interface may be attributed to diffusive fluxes or advective transport (e.g. by tidal pumping) of oxidants from the overlying water into the sediment profile. The pyrite content at both sites generally increased with depth below the sediment–water interface. At site W1, pyrite-S increased from 63 lmol g1 in the 0–2 cm depth interval to 251 lmol g1 in the 18–22 cm depth interval (Fig. 2). Considerably greater concentrations of pyrite were present at site W2, with pyrite-S ranging from 312 to 669 lmol g1. Greater amounts of pyrite at depth are reflected in decreasing AVS to pyrite-S ratios (Fig. 2). These trends in AVS to pyrite-S ratios are likely to be caused by formation of pyrite from meta-stable FeS precursors (Schoonen and Barnes, 1991). This process would lead to depletion of AVS in deeper, older sediments and a co-existing relative enrichment of pyrite-S. P Sequestration of pore-water S(-II)(aq) by precipitation of solid-phase Fe–S species is partly controlled by supply of Fe (Boesen P and Postma, 1988). The moderately high porewater S(-II)(aq) and very low pore-water Fe concentrations suggest that formation of solid-phase Fe–S species may be Fe-limited in the sediments examined here. The total Fe contents spanned 254–409 lmol g1 at site W1 and 305–790 lmol g1 at site W2 (Fig. 3). However, a large

Fig. 2. Abundance of sulfur fractions in sediment profiles from sites W1 and W2. Data points represent the mean ± 1 standard deviation of duplicate determinations (in cases where error bars are not presented, the magnitude of ±1 standard deviation is less than the symbol size).

pore-water Fe(II) to low concentrations (i.e. <1 lmol l1) at near-neutral pH. Even lower Fe(II) concentrations would exist if more crystalline Fe–S phases, such P as mackinawite, greigite or pyrite, controlled Fe(II) and S(-II)(aq) solubility (Morse et al., 1987). The very low, but detectable pore-water Fe(II) concentrations in the near-surface sediments (0–4 cm) are probably due to diffusion from a very thin, sub-oxic layer at the sediment–water interface (Teasdale et al., 2003). Maximum AVS concentrations of 129 lmol g1 and 32.7 lmol g1 were present in near-surface sediment at sites W1 and W2, respectively. These concentrations are comparable to those presented in Burton et al. (2005), who found AVS ranging up to 113 lmol g1 in surface sediments from nearby sites. There was a decrease in AVS concentrations with depth below the sediment–water interface, with 1 10 lmol P g observed below 14 cm at both sites. Porewater S(-II)(aq) comprised between 1% and 3% of AVS near the sediment–water interface, indicating that nearsurfacePAVS mostly consisted of solid-phase S(-II). Porewater S(-II)(aq) comprised a more substantial proportion (6–8%) of AVS at depths greater than 10 cm below the sediment–water interface at site B2. Elemental S was most abundant (27 lmol g1 at site W1 and 33 lmol g1 at site W2) in the 0–2 cm surface layer (Fig. 2). These S08 ðsÞ concentrations indicate oxidation of reduced inorganic S (RIS) near the sediment surface (Troelsen and Jørgensen, 1982). Lower S08 ðsÞ levels at

Fig. 3. Abundance of total Fe and individual Fe fractions in sediment profiles from sites W1 and W2 (FeS–Fe and FeS2–Fe were determined stoichiometrically from RIS results; Reactive Fe was defined by 1 M HCl extraction; ‘‘Reactive Fe (other)’’ refers to non-sulfidised reactive Fe, calculated by subtracting FeS–Fe from 1 M HCl-extractable Fe; Residual Fe was defined as total Fe minus the sum of reactive Fe and FeS2–Fe). Data points represent the mean ± 1 standard deviation of duplicate determinations.

E.D. Burton et al. / Chemosphere 64 (2006) 1421–1428

proportion of total Fe (generally >40%) was occluded as non-pyrite, non-reactive (or very slowly reactive) phases (‘‘residual’’ Fe in Fig. 3). This is probably Fe bound within the lattice structure of crystalline silicates, which are poorly P reactive with pore-water S(-II)(aq) (Canfield et al., 1992). The effect of reactive Fe availability on sulfide accumulation may be described by the degree of sulfidisation (DOS; Boesen and Postma, 1988): DOS ¼ 100 

AVSFe þ pyriteFe reactiveFe þ pyriteFe

ð1Þ

where AVSFe and pyriteFe denote Fe associated with solidphase AVS (FeS) and pyrite, respectively; and reactiveFe represents 1 M HCl-extractable Fe. The 1 M HCl extraction recovers reactive non-sulfide phases as well as FeS (thereby inherently including AVSFe). The DOS value includes both reactiveFe as well as less labile pyriteFe. As such, it is useful to also determine the reactive sulfidisation of Fe associated with non-pyrite phases. This may be determined by calculation of a parameter, which we hereby term the degree of reactive sulfidisation (DORS): DORS ¼ 100 

AVSFe reactiveFe

ð2Þ

The DOS value was close to 100% in near-surface sediment at both sites (Fig. 4). At site W2, the DOS value remained high (>85%) throughout the 0–22 cm depth interval examined here. The DOS value decreased with depth from the sediment surface to the 8–10 cm depth interval at site W1, before stabilizing at 80–90% at greater depth (Fig. 4). These generally high DOS values indicate that large proportions of the sum of reactive- and pyrite-Fe are associated with sulfide phases. The DORS values were variable, spanning 21–100% at site W1 and 19–63% at site W2 (Fig. 4). Near-surface sediment at site W1 exhibited DORS values close to 100%, indicating that almost all reactive Fe was present as FeS (assuming that solid-phase AVS has a 1:1 Fe:S stoichiometry). In contrast, the low DORS values (20%) in the lower depth intervals at both sites indicate that up to 80% of reactive Fe was not sulfidised.

Degree of Sulfidisation (%)

0

0 25 50 75 100

Depth (cm)

(W1)

0 25 50 75 100 (W2)

5 10 15 20

Fig. 4. Degree of sulfidisation (DOS, open circles) and degree of reactive sulfidisation (DORS, filled circles) in sediment profiles from sites W1 and W2. Data points are the mean ± 1 standard deviation of duplicate determinations.

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The mostly low DORS values show an apparent excess of reactive Fe in the majority of depth intervals examined (except 0–6 cm at site W1; Fig. 4). This is unexpected because theP high DOS values (Fig. 4) and moderately high pore-water S(-II)(aq) (Fig. 2) suggest that supply of reactive Fe may limit formation of Fe–S minerals. Slow rates of reactive Fe sulfidisation along with efficient conversion of AVSFe to pyriteFe (indicated by low AVS to pyrite-S ratios) may cause these low DORS and high DOS values. Slow sulfidisation of reactive Fe may be due to ‘‘armouring’’ of the particle surface by sulfide phases, thereby occluding the interior Fe (Canfield et al., 1992). 3.3. Trace element extractability Total trace element concentrations were in the range of 2.1–21.4 lg g1 for As, 0.08–0.63 lg g1 for Cd, 34.2–63.8 lg g1 for Cr, 13.2–83.4 lg g1 for Cu, 16.7–26.1 lg g1 for Ni, 15.0–94.5 lg g1 for Pb, and 73.4–246 lg g1 for Zn (Fig. 5). Total concentrations of As, Cu, Ni, Pb and Zn in some depth intervals at sites W1 and W2 exceeded the Australian sediment quality trigger values (20 lg g1 for As, 65 lg g1 for Cu, 21 lg g1 for Ni, 50 lg g1 for Pb, and 200 lg g1 for Zn; ANZECC/ARMCANZ, 2000). At these concentrations, trace elements may represent a threat to sediment and water quality and associated ecosystem functions. Measurement of metal reactivity, rather than total concentrations, provides a more accurate description of sediment quality (Chapman et al., 1998; ANZECC/ARMCANZ, 2000). Extractions with cold, dilute HCl (0.5–6 M) have been widely used as an indication of trace element reactivity and hence availability (Di Toro et al., 1992; Cooper and Morse, 1998a; McCready et al., 2003; Burton et al., 2005). In the present study, the reactive trace element fraction is defined by extraction with 1 M HCl for 1 h at room temperature. For Cd, Pb and Zn, 1 M HCl-extractable concentrations were comparable to total concentrations (Fig. 5). The percentage of total Pb recovered by 1 M HCl (i.e. extractability) was not significantly correlated with the total Pb concentrations (Table 2). In contrast, the extractability of Cd and Zn was positively correlated with the total concentration of these elements (Table 2). This indicates that, although Cd and Zn extractability is generally high (mostly >70% recovered by 1 M HCl; Fig. 5), these elements become increasingly reactive as the total concentration increases. This may be attributed to saturation of high affinity sorption sites at higher total Cd and Zn concentrations, thereby enhancing sorption to more reactive sites. Extractable (1 M HCl) concentrations of As, Cr, Cu and Ni were low (<6 lg g1) compared to the corresponding total concentrations (Fig. 5). Generally, the reactive fraction was <20% of the total concentration of these elements at sites W1 and W2. The extractability of Cr, Cu and Ni was not positively correlated (P > 0.05) with their respective total concentrations (Table 2). This shows that the

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Site W1 Total and 1M HCl-extractable trace elements (μg g-1) 0

20

40

60

0

40

80

120

0

70

140 210

0

10

20

30

0

0.3

0.6

0.9

0

3

6

0

9

20

40

60

Depth (cm)

0 5 10 15

(Cu)

20

(Pb)

(Cr)

(As)

(Cd)

(Ni)

(Zn)

Site W2 0

30

60

90

0

40

80

120

0

100 200 300

0

10

20

30

0

0.3

0.6

0.9

0

7

14

0

21

20

40

60

Depth (cm)

0 5 10 15 20

(Cu)

(Pb)

(Cr)

(As)

(Cd)

(Ni)

(Zn)

Fig. 5. Total (filled circles) and 1 M HCl-extractable trace element (open circles) concentrations in sediment profiles from sites W1 and W2. Precision associated with each data point is within 10% (except for 1 M HCl-extractable Cu, in which case the precision is within 25%).

Table 2 Pearson correlation coefficients for relationships between trace element extractabilitya and selected sediment properties (n = 16) As 1

Total concentration (lg g ) AVS (lmol g1) Pyrite-S (lmol g1) AVS/pyrite-S Total C (%) Reactive Fe (lmol g1)

Cd **

0.707 0.509* 0.800** 0.667* 0.343 0.311

*

0.576 0.095 0.354 0.114 0.035 0.238

Cr

Cu

Ni

Pb

Zn

0.330 0.159 0.584* 0.052 0.022 0.328

0.137 0.303 0.736** 0.459 0.317 0.135

0.041 0.353 0.268 0.162 0.132 0.543*

0.265 0.213 0.360 0.351 0.187 0.170

0.759** 0.021 0.398 0.279 0.119 0.058

Significant correlations with P < 0.01 are shown in bold. * P < 0.05. ** P < 0.01. a Extractability (%) = 100 * reactive/total.

total concentration provides a poor indication of the reactivity of Cr, Cu and Ni. Arsenic extractability was negatively correlated with total As (Table 2), demonstrating that As reactivity decreased as the total concentration became greater. Many trace metals, particularly Cd, Cu, Ni, Pb and Zn, form very poorly soluble sulfide minerals that are stable under reducing conditions (Chapman et al., 1998). An excess of AVS over reactive trace metals is expected to cause very low trace metal pore-water concentrations due to precipitation of metal sulfides. The measured AVS concentrations exceed the corresponding sum of reactive Cd, Cu, Ni, Pb and Zn concentrations (in lmol g1) by 6–87 times for site W1 and 6–11 times for site W2. The excess AVS over reactive trace metals indicates that Cd, Cu, Ni, Pb and Zn bioavailability may be low (Di Toro et al., 1992; Ankley et al., 1994; Simpson et al., 2002). The AVS to pyrite-S ratios indicate progressive replacement of AVS by pyrite with increasing depth below the sediment–water interface (Fig. 2). The pyritisation of AVS may be expected to also result in pyritisation of trace elements associated with AVS (Belzile and Lebel, 1986;

Huerta-Diaz and Morse, 1992). This would lead to a transition from trace element association with reactive AVS in near-surface sediment to sequestration in pyrite with increasing sediment depth. This process of trace element adsorption to, and co-precipitation within pyrite would cause a corresponding decrease in 1 M HCl-extractability with depth. The near-total recovery of Cd and Pb by extraction with 1 M HCl (Fig. 5) suggests that these metals were not substantially bound by pyrite during pyritisation of AVS. Zinc was also readily extracted, with generally >70% of total Zn recovered by extraction with 1 M HCl for the sediments examined in this study (Fig. 5). This is consistent with Huerta-Diaz and Morse (1992), who found that Cd, Pb and Zn were not substantially pyritised in anoxic marine sediments. Zinc, Cd and Pb have filled d-orbitals resulting in no ligand field stabilisation, whereas the d-orbitals in Fe(II) are partially filled. As a result, Cd, Pb and Zn have fast rates of water exchange and react to form CdS, PbS and ZnS phases faster than Fe2+ forms FeS (Morse and Luther, 1999). In sulfidic sediments, this causes Cd, Pb and Zn to form distinct sulfide phases, thereby reducing

E.D. Burton et al. / Chemosphere 64 (2006) 1421–1428

their tendency to be incorporated into pyrite during pyritisation of FeS. The average 1 M HCl-extractability observed for Cu and Ni in the sediments examined here was 6% and 12%, respectively (Fig. 5). Adsorption or co-precipitation of Cu and Ni with pyrite may account for this low extractability (Huerta-Diaz and Morse, 1992). However, low Cu and Ni extractability is also consistent with the recovery of their pure sulfide minerals in cold HCl. For example, Cooper and Morse (1998b) found that only 1% of NiS2, 23% of NiS, 12% of CuS and 18% of Cu2S was recovered by HCl extraction. Copper may also form several poorly extractable Cu–Fe–S phases, as well as solid-solutions with pyrite where a distinct Cu–Fe–S phase is not evident (Morse and Luther, 1999). Negative correlation between Cu extractability and the pyrite-S concentrations tentatively support the hypothesis that Cu may be associated with pyrite at sites W1 and W2 (Table 2). However, future research is needed to determine if the low observed Cu and Ni extractability is caused by (1) presence of pure Cu and Ni sulfide minerals, (2) association with pyrite during pyritisation of AVS, or (3) binding by resistant non-sulfidic phases. Arsenic extractability was consistently low (Fig. 5), averaging only 19% of the total As concentration. There was a strong negative correlation between As extractability and the pyrite-S concentrations (Table 2). Arsenic extractability was also positively correlated with the observed AVS to pyrite-S ratios (Table 2). This is consistent with As being occluded by adsorption/co-precipitation with pyrite or by precipitation of arsenopyrite with increasing depth below the sediment–water interface. The results suggest that As may be bound by pyrite or arsenopyrite as these phases replace AVS during diagenesis. On average, only 9% of total Cr in the sediments examined here was recovered by 1 M HCl extraction (Fig. 5). Chromium extractability was positively correlated with pyrite-S concentrations, which suggests that Cr availability was not reduced by association with pyrite. Morse and Luther (1999) state that structural disparity between the primary Cr–Fe–S mineral (daubre´elite, Cr3FeS4) and pyrite means that Cr does not co-precipitate with pyrite. Furthermore, Cr(III) is kinetically inert to reaction with porewater sulfide (Morse and Luther, 1999), thereby retarding sulfidisation of Cr in sulfidic sediments. O’day et al. (2000) found that Cr was ligated by O Patoms in sediments containing high levels of pore-water S(-II)(aq), reflecting association with oxide, carbonate or silicate minerals rather than binding by sulfides. Therefore, the weak positive correlation between Cr extractability and pyrite-S concentrations presented in Table 2 probably reflects co-occurring diagenetic processes. For example, sulfidisation of crystalline Fe oxides during diagenesis would release poorly reactive oxide-bound Cr. This sulfidisation process would also co-occur with pyritisation of AVS, and may therefore cause positive correlation between Cr extractability and pyrite-S (Table 2).

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4. Conclusions P The moderately high pore-water S(-II)(aq) and AVS concentrations suggest that trace elements in the sediments described here are likely to be bound as sulfide minerals. The AVS to pyrite-S ratios at the study sites reflect replacement of AVS by pyrite with increasing depth. Cadmium, Pb and Zn extractability indicated that these elements are not significantly sequestered during pyritisation of AVS. Trends in As and Cu extractability suggest that these elements may be sequestered during pyrite formation. Nickel extractability was low, possibly due to formation of pure Ni–S phases or sequestration by pyrite. The results suggest that Cr may become more reactive during diagenesis in sulfidic sediments. In conclusion, replacement of AVS by pyrite during diagenesis may enhance the reactivity of Cd, Cr, Pb and Zn, whereas As, Cu and possibly Ni may be rendered non-reactive. Acknowledgements This work forms part of a project examining ‘‘contemporary sulfur biomineralisation in acid sulfate soil landscapes’’, which was funded by the Australian Research Council (DP0453280). The authors thank Dane Lamb for assistance with sediment collection. References Allen, H.E., Fu, G., Deng, B., 1993. Analysis of acid-volatile sulfide (AVS) and simultaneously extracted metals (SEM) for estimation of potential toxicity in sediment. Environ. Toxicol. Chem. 12, 1441–1453. Ankley, G.T., Thomas, N.A., Di Toro, D.M., Hansen, D.J., Mahony, D.J., Berry, W.J., Swartz, R.C., Hoke, R.A., Garrison, A.W., Allen, H.E., Zarba, C.S., 1994. Assessing potential bioavailability of metals in sediments: a proposed approach. Environ. Manage. 18, 331–337. ANZECC/ARMCANZ, 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality. Australian and New Zealand Environment and Conservation Council, and Agriculture and Resource Management Council of Australia and New Zealand. APHA, 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association–American Water Works Association, Baltimore, USA. Belzile, N., Lebel, J., 1986. Capture of As by pyrite in near-shore marine sediments. Chem. Geol. 54, 279–281. Boesen, C., Postma, D., 1988. Pyrite formation in anoxic sediments of the Baltic. Am. J. Sci. 288, 575–603. Burton, E.D., Phillips, I.R., Hawker, D.W., 2005. Reactive sulfide relationships with trace metal extractability in sediments from southern Moreton Bay, Australia. Mar. Pollut. Bull. 50, 589–608. Canfield, D.E., 1989. Reactive iron in marine sediments. Geochim. Cosmochim. Acta 53, 619–632. Canfield, D.E., Raiswell, R., Bottrell, S., 1992. The reactivity of sedimentary iron minerals towards sulfide. Am. J. Sci. 292, 659–688. Chapman, P.M., Wang, F., Janssen, C., Persoone, G., Allen, H.E., 1998. Ecotoxicity of metals in aquatic sediments: binding and release, bioavailability, risk assessment, and remediation. Can. J. Fisher. Aquat. Sci. 55, 2221–2243. Clark, M.W., 1998. Management implications of metal transfer pathways from a refuse tip to mangrove sediments. Sci. Total Environ. 222, 17– 34.

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