Suboxic, early diagenetic processes in surficial sediments near a deepwater ocean outfall, Sydney, Australia

ELSEVIER

Journal of Geochemical Exploration 64 (1998) 1–17

Suboxic, early diagenetic processes in surficial sediments near a deepwater ocean outfall, Sydney, Australia C. Matthai a,Ł , G.F. Birch a , R. Szymczak b b

a Environmental Geology Group, School of Geosciences, University of Sydney, NSW 2006, Sydney, Australia Australian Nuclear Science and Technology Organisation, Environment Division, PMB 1, Menai, NSW 2234, Australia

Accepted 12 May 1998

Abstract Triplicate boxcores collected from three sample locations, near the Malabar deepwater ocean outfall on the continental margin adjacent to Sydney, were analysed on a total and a size-normalized basis (<63 µm fraction) to establish the extent of heavy metal enrichment. No heavy metal enrichment near the outfall is observable in total sediment samples, but an enrichment in Ag, Cu, Pb and Zn is detectable in the sediment fine fraction and in weak acid leachable bulk sediment concentrations. Trace metal profiles for the fine fraction indicate that chemical remobilization, and physical sediment mixing and resuspension influence downcore trace metal distributions. Ratios of solid and dissolved-phase porewater Fe, Mn and Cu indicate that oxidation of sewage organic matter and suboxic sulphide reduction influence dissolved-phase trace metal concentrations in sediments near the outfall. Suboxic, early diagenesis results in an increased dissolved-phase benthic flux of Cu near the outfall, with organic matter reduction likely to be the major diagenetic process.  1998 Elsevier Science B.V. All rights reserved. Keywords: continental shelf; sediment; trace metals; deepwater ocean outfall; sewage; sediment porewater

1. Introduction Environmental effects of marine sewage discharge have recently received considerable attention in Australia (Fagan et al., 1992; Scanes and Philip, 1995; EPA, 1996; Sydney Water, 1996) and other countries (Webster and Ridgway, 1994; Chapman et al., 1996). The main environmental concerns about marine sewage discharge are the accumulation of organic matter and trace metals in sediment, water and biota near the discharge. Ł Corresponding

@es.su.oz.au

author. Fax: C61.2.93510184; E-mail: carsten

Ninety-four percent of the sewage effluent generated in the Sydney metropolitan area (population approximately 3.5 million) currently only receives primary treatment and is discharged into the ocean via three deepwater ocean outfalls at a rate of 915 Ml d 1 (Sydney Water, 1996). The Malabar sewage treatment plant is the largest in Sydney, treating waste water from a catchment area of 500 km2 , about a quarter of the flow being from commercial and industrial sources. Almost half (475 Ml d 1 ) of Sydney’s sewage effluent is discharged via a 3.5-km-long deepwater (about 80 m depth) ocean outfall on the middle shelf off Malabar along a 720 m diffuser zone.

c 1998 Elsevier Science B.V. All rights reserved. 0375-6742/98/$19.00 PII: S 0 3 7 5 - 6 7 4 2 ( 9 8 ) 0 0 0 1 8 - 1

2

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

Benthic biota, sediments and water quality in the vicinity of the outfall have been assessed and monitored extensively before and after commissioning in September, 1990 as part of the Sydney Deepwater Outfalls Environmental Monitoring Program (Fagan et al., 1992; Scanes and Philip, 1995; Schneider and Davey, 1995; EPA, 1996; Matthai and Birch, 1996; Sydney Water, 1996). No adverse effects on water and sediment quality have been detected, suggesting that most of sewage effluent is efficiently dispersed on this high energy shelf. However, the use of bulk sediment trace metal concentration data obtained during monitoring of the Malabar outfall (EPA, 1996), to record anthropogenic impacts is limited by grain size-controlled trace metal variability and a lack of knowledge regarding chemical speciation of the trace metals. Recent investigations of surficial sediments on the central NSW continental margin have shown a regional anthropogenic enrichment of trace metals associated with the fine fraction (<63 µm) adjacent to Sydney, attributable to contributions from a multitude of sources (Matthai and Birch, 1995; Birch, 1996). The relationship between the dissolved and particulate trace metal fractions has been used elsewhere to identify the nature and extent of chemical reactions that take place between the sediments and interstitial water (Balls, 1989; Calmano et al., 1990; O’Reilly Wiese et al., 1997). Porewater and overlying water Fe, Mn and Cu concentrations are used in this study to assess the diffusive flux of trace metal across the sediment–water interface when solute concentrations in the sediment are higher than in the overlying bottom water (see also Skowronek et al., 1994; Petersen et al., 1995). The degree to which sedimentary trace metals are bioavailable and the extent of bioaccumulation is commonly related to dissolved-phase trace metal concentrations (Bryan, 1984; O’Reilly Wiese et al., 1997). However, recent work has shown that the toxicity of polluted sediment to benthic fauna may be more closely related to the contaminant concentration in interstitial water than to the bulk sediment concentration (di Toro et al., 1992; Schults et al., 1992). In contrast to the stable concentrations in the solid phase, the porewater concentration gradients commonly reflect sediment redox conditions (Klinkhammer et al., 1982).

Concentrations of free trace metal ions in porewater depend on early diagenetic processes in the sediments and are mainly driven by organic matter fluxes and mineralization (Klinkhammer, 1980; Petersen et al., 1995). Bioreactive trace metals, such as Cu, have been shown to be diagenetically remobilized by aerobic decomposition of biogenic or sewage-derived organic matter, and thus display a nutrient-like behaviour in the marine environment (Klinkhammer et al., 1982; Widerlund, 1996). For sedimentation rates above about 0.1 cm y 1 , sulphate reduction and oxic respiration are equally important in oxidizing organic matter, whereas oxic respiration decomposes about 100–1000 times more organic carbon than sulphate reduction in sedimentary environments with sedimentation rates of 10 3 cm y 1 (Canfield, 1989). Sedimentation rates on the Sydney middle shelf exceed 0.1 to 0.2 cm y 1 (C. Matthai, unpubl. data) and hence, it is likely that sulphate reduction is at least as important as oxic respiration in the decomposition of organic matter.

2. Study area The continental shelf adjacent to Sydney is characterised by unconsolidated sediments, which gradually fine offshore, and by extensive bedrock (Roy, 1985). A tripartite subdivision of the shelf into inner, middle and outer zones is associated with three shore-parallel Quaternary shelf lithofacies (Davies, 1979; Roy, 1980): (a) inner shelf, medium- to coarse-grained, yellowish-orange, subangular quartzose sands (water depth <60 m); (b) middle shelf, fine- to medium-grained, olive-green, slightly calcareous, angular quartzic muds to muddy sands (water depth 60–120 m); (c) outer shelf, coarsegrained, reddish-brown, rounded, calcareous, ironstained gravelly sands (water depth >120 m). The oceanographic conditions on the Sydney continental margin are dominated by the southwardflowing East Australian Current, a complex system of migrating eddies (Gordon and Hoffman, 1986). Strong wind-induced bottom currents of up to 1 m s 1 flow in a general southerly direction throughout the year (EPA, 1995).

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

3. Materials and methods 3.1. Sample collection Three sites on the middle shelf in water depths of 80 m (MBC1), 84 m (MBC2) and 89 m (MBC3) were sampled. MBC1 is approximately 180 m south of the outfall and MBC2 is 1800 m south of MBC1; these sites were chosen after a pilot survey showed that sewage impact on surficial sediment trace metal concentrations was to the south of the outfall, in concordance with the major current flow direction in this area (Matthai and Birch, 1996). MBC3, 8300 m southwest of the outfall, is a reference site where trace metal concentrations in the fine fraction of the sediment were found to approach regional baseline values (Matthai and Birch, 1996) (Fig. 1).

3

Each site was sampled in triplicate, using a Soutar boxcorer to obtain an undisturbed surficial sediment sample. Differential GPS and the deployment of a mooring buoy ensured triplicate samples were taken within a 50-m radius. Three subcores of each boxcore were taken after overlying water was drained to about 10 cm depth. Subcores were removed, capped, and transferred, with an overlying water sample, to an insulated box kept at in-situ sediment temperatures. The three subcores collected from each boxcore were used for the determination of: (a) porewater Fe, Mn, Cu and 0.05 M EDTA-extractable bulk sediment Ag, Co, Cu, Fe, Mn, Ni, Pb, Zn; (b) fine and coarse fraction Ag, Co, Cu, Fe, Mn, Ni, Pb, Zn, TOC, TIC; and (c) sediment porosity.

Fig. 1. Sample sites.

4

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

3.2. Sample pretreatment One subcore from each site was extruded at 1 cm intervals (0–1 cm; 1–2 cm; 2–3 cm; 4–5 cm; 9–10 cm; 14–15 cm; 19–20 cm) under a nitrogen atmosphere inside an I2 R inflatable glove chamber and an overlying water sample was collected at the same time. Porewater was extracted from three cores (MBC1A, MBC2A and MBC3A) by centrifugation under nitrogen at 4ºC (4500 rpm for 20 min) and both overlying water and porewaters were filtered through 0.4 µm acid-washed Poretics polycarbonate membrane filters. The filtered overlying waters and porewaters were transferred to 20-ml polyethylene scintillation vials and acidified to ¾pH 1 with Merck Suprapur HNO3 . Procedural blank 18.2 M Ohm.cm Milli-Q water samples were processed at the time of porewater extraction to assess laboratory contamination. All sample storage and handling equipment was washed in detergent, acid cleaned in a 10% HNO3 bath for 24 h, rinsed with 18.2 M Ohm.cm Milli-Q water and dried in a Gelman laminar flow clean air cabinet. Weak acid leaching was performed in 50-ml Falcon polypropylene centrifuge tubes using 0.05 M ethylenediaminetetraacetic acid (EDTA) as an extractant (Ying et al., 1992). The extractant: bulk sediment ratio of 25 : 1 (v=w) is sufficient for acid neutralization by sediment decarbonisation to be negligible (Ying et al., 1992). EDTA leaching was promoted by shaking the samples for 2 h using a circular motion stirrer set at 60 rpm. Subsequent centrifugation at 4500 rpm at 4ºC for 20 min and a transfer of the extractant solutions from the polypropylene centrifuge tubes into 20-ml polyethylene scintillation vials completed the EDTA leaching procedure. A second subcore from each of the three triplicate samples from each site was sliced at 1-cm intervals using the same sampling intervals as described previously and size-normalized by wet-sieving through a 63-µm nylon mesh to separate fine and coarse fractions. No gravel was present in the coarse fraction. The water in which the fine fraction settled contained no measurable trace metal concentrations, exemplifying a negligible trace metal loss from the solid phases into the dissolved phase (C. Matthai, unpubl. data). Both mud and sand fractions were

dried at 60ºC for 48 h, weighed, and homogenized by grinding in an agate mortar and pestle. Both sediment mud and sand fractions were gravimetrically corrected for salt by adding deionized water to a preweighed subsample, settling the suspension, decanting the overlying water, and reweighing the dried, salt-free sample. About 0.5 g of each fine and 1 g of each coarse fraction sample were weighed into separate 50-ml Pyrex borosilicate test tubes, and 10 ml of a 2 : 1 mixture of concentrated HClO4 –HNO3 acid was added to each test tube (Irvine, 1980; Birch and Davey, 1995). The test tubes were heated to 160ºC for 12 h or until acid evaporation left the samples close to dryness. After cooling, 30 ml of ultrapure water was added and the test tubes were shaken in a rotation stirrer to homogenize the extraction solution. After settling, the extraction solutions were transferred to 20-ml polyethylene scintillation vials and stored at 4ºC. Total sediment trace metal concentrations were determined by recombining the separately determined concentrations of the sediment fine and coarse fractions according to Eq. 1: CBS D CM ð FM C CS ð FS

(1)

where CM and CS are the trace metal concentrations in the mud and sand fractions (µg g 1 ), and FM and FS are the respective proportions of mud and sand. Independent studies have shown this to accurately represent the total sediment heavy metal concentration (Birch and Davey, 1995; Birch, 1996). 3.3. Analytical methods 3.3.1. Porewater and overlying water analysis Porewater and overlying water samples from cores BCM1A, BCM2A and BCM3A were analyzed in triplicate for Fe, Mn and Cu, with an analytical variance of <30% relative standard deviation (RSD) in all samples. Procedural blanks were below the detection limit (Fe: 16 nM; Mn: 1 nM; Cu: 5 nM) for all metals. Analyses were performed in a Class 350 Clean Room laboratory by direct injection on a Perkin-Elmer SIMAA 6000 graphite furnace simultaneous multi-element atomic absorption spectrophotometer (GFAAS) equipped with Zeeman background correction. Extended charring steps dur-

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17 Table 1 Uncertainties in the porewater metal data (nM) Core

Fe (nmol l 1 )

Mn (nmol l 1 )

Cu (nmol l 1 )

MBC1A MBC2A MBC3A

34 10 232

5.5 2.2 4.0

1.2 1.4 1.0

5

Kern and Westrich, 1995). Carbonate carbon of fine fractions was determined by HCl sample decarbonization in a Karbonat Bombe with a precision of š1% for the measured concentration range (Birch, 1981).

4. Results and discussion ing the analysis of each metal minimized matrix effects. Samples were diluted by up to 50 : 1 for the Fe and Mn analyses. The precision for each element is the mean of the standard deviations for the triplicate analyses of each sample (Table 1). 3.3.2. Sediment analysis The 0.05 M EDTA-leachable bulk sediment fraction and the mud and sand strong acid extractable fractions of the sediments were analyzed for Ag, Co, Cu, Fe, Mn, Ni, Pb and Zn using a PerkinElmer 3100 instrument. The detection limits for the analyses are given in Table 2. The reproducibility of the strong acid extraction was tested by digesting seven samples in duplicate and comparing the trace metal concentrations of the duplicates. The variation in the duplicate analyses was <10% RSD for all elements (Ag: <6%; Co: <6%; Cu: <3%; Fe: <10%; Mn: <5%; Ni: <7%; Pb: <4%; Zn: <8%), as were replicate analyses .n D 7/ of internal laboratory standard reference material ILS-2. Total organic carbon (TOC) in fine and coarse fractions was approximated by gravimetric loss on ignition at 550ºC for 2 h (Dankers and Laane, 1983; Table 2 Detection limits for fine and coarse fraction extractions and bulk sediment leachings (µg g 1 ) Element

Ag Co Cu Fe Mn Ni Pb Zn

Fine fraction extraction (µg g 1 )

Coarse fraction extraction (µg g 1 )

Bulk sediment leaching (µg g 1 )

0.6 1.2 1.2 12 12 1.8 3 6

0.3 0.6 0.6 6 6 0.9 1.5 3

0.1 0.3 0.3 2.5 2.5 0.4 0.6 1.3

4.1. Sediment physicochemical parameters To facilitate interpretation of the dissolved-phase fluxes at each location, it is necessary to describe the physicochemical characteristics of the sediments. All downcore heavy metal data are summarized as means of triplicate samples from each location š1 standard deviation (SD) (Table 3). The small-scale spatial variability (SSSV) of mud content is highest at MBC2, with a mean relative standard deviation (RSD) of 37% in the surficial 10 cm, compared to 14% at MBC1 and 19% at MBC3. Surficial sediment textures range from sandy (up to 14% mud) at MBC1, to muddy-sand at MBC2 (up to 23% mud) and MBC3 (up to 24% mud) (Fig. 2a). Mean sediment porosity decreases downcore at all three locations and is lowest at MBC1 (<50%), compared to values of 68% and 62% at MBC2 and MBC3, respectively (Fig. 2b). Mean downcore TOC content in the fine fraction is essentially constant at MBC2 and MBC3 (15– 17%) and decreases slightly from 19% at the surface to 14% at 9 to 10 cm depth at MBC1 (Fig. 2c). At MBC1 the TOC content increases downcore from Table 3 Mean fine fraction trace metal concentrations (µg g 1 , Fe in %) for five sediment samples (0–10 cm) from each of the triplicate cores at MBC1, MBC2 and MBC3 (n D 15; numbers in brackets indicate 1 SD) Element

MBC1

MBC2

MBC3

Ag Co Cu Fe Mn Ni Pb Zn

4.5 (0.3) 8.1 (0.7) 73 (6) 3.94 (0.45) 122 (6) 28.4 (1.5) 58 (6) 163 (12)

3.9 (0.2) 8.1 (0.6) 59 (3) 3.63 (0.26) 123 (7) 28.8 (1.5) 51 (4) 144 (8)

3.4 (0.2) 7.9 (0.4) 48 (3) 3.80 (0.26) 127 (9) 26.6 (0.9) 45 (3) 128 (6)

6

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

a mean of 3.4% at the surface to 4.7% at 10 cm depth, whereas at MBC2 and MBC3 it decreases substantially downcore (Fig. 2d). The carbonate content of the fine fraction is almost constant at all three locations, with mean concentrations ranging from 24.7% to 26.0% at the surface to 23.8% to 24.2% at 9 to 10 cm depth (Fig. 2e). The total sediment TOC content correlates strongly with mud content (R 2 D 0:85; p < 0:05; n D 50) and with total sediment trace metal concentrations (R 2 > 0:57; p < 0:05; n D 50). Differences in downcore TOC concentrations at the three locations are therefore mainly texturally controlled and do not reflect current organic matter input. There is no correlation between fine fraction TOC content and mud content. Slightly elevated fine fraction TOC concentrations in the sediment at MBC1 are most likely the result of increased particulate organic matter flux to the sediments from the adjacent deepwater ocean outfall. 4.2. Total and size-normalized sediment trace metals Downcore total sediment trace metal profiles are similar at each sample location. MBC1 total sediment trace metal concentrations increase downcore to a depth of approximately 5 cm, below which concentrations remain constant. Total sediment trace metal SSSV for MBC1 is low, with a mean RSD in the surficial 10 cm of sediment of 5–13%. At MBC2, total sediment trace metal concentrations are generally higher in the upper 2–3 cm than at depth, with levels decreasing downcore to concentrations below that at MBC1. SSSV of total sediment trace metals is highest at MBC2, with a RSD in the top 10 cm ranging from 14 to 32%, indicating high textural variability. Bulk sediment trace metal concentrations are generally highest at MBC3 (the control site), with profiles displaying a slight downcore decrease for all elements to a depth of 4–5 cm, below which concentrations remain nearly constant (Fig. 3). The mean total sediment trace metal RSD in the top 15 cm is between 8 and 12%.

7

Overall, total sediment trace metal concentrations in surficial sediments are low at all three sample locations and an anthropogenic trace metal enrichment from the sewage outfall is not apparent. The relationship between mud content and total sediment trace metal concentrations, however, reveals that a trace metal enrichment could be masked by textural variability. Correlations between mud concentrations and total sediment heavy metal concentrations have R 2 values of 0.40 (Ag), 0.62 (Co), 0.82 (Cu), 0.84 (Fe), 0.45 (Mn), 0.91 (Ni), 0.62 (Pb) and 0.88 (Zn) ( p < 0:05; n D 50). Size-normalization (<63 µm) largely overcomes texturally induced variability in trace metal data and clearly reveals an anthropogenic contribution to trace metal loads. The RSD for fine fraction trace metals is below 6.5% at all three locations, with the exception of Fe (9.4%) and Co (8.4%) at MBC1. Fine fraction Cu, Pb, Zn and Ag concentrations decline away from the outfall, with higher mean concentrations in the surficial 10 cm at MBC1 than at the other two locations. Ag in particular has previously shown to be a good sewage tracer in coastal sediments (Ravizza and Bothner, 1996). Mean fine fraction concentrations of Fe, Mn, Co and Ni, on the other hand, remain nearly constant at the three locations, indicating lower anthropogenic contributions compared to Ag, Cu, Pb and Zn (Fig. 4 and Table 4). Sediments at MBC1 display a slight downcore increase in fine fraction trace metal concentrations to a depth of about 5 cm for Cu and Ag and to 10 cm for Pb and Zn. This increase is unlikely to result from a downcore decrease in grain size within the fine fraction, or due to dilution by fine sediment poor in trace metals, such as calcium carbonate (micrite), which would affect the concentrations of all elements to the same extent. This is not the case for fine fraction Fe, Mn and Co, which display little downcore variation in concentration (Fig. 4). Furthermore, the minor or slight enrichment of fine fraction carbonate in the surficial 2–3 cm (Fig. 2e) is not sufficient to explain the observed decrease in the concentration of fine fraction trace metals near the surface. A chemical remobilization of Ag, Cu,

Fig. 2. Mud fraction (a); porosity (b); fine fraction TOC (c); total sediment TOC (d) and fine fraction carbonate (e) profiles for sediments at MBC1, MBC2 and MBC3. All data are means of triplicate cores, except for MBC2 (14–15 cm) and MBC3 (19–20 cm), where n D 1.

8

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

Fig. 3. Bulk sediment trace metal profiles for triplicate cores at MBC1, MBC2 and MBC3. All data are means of triplicate cores, except for MBC2 (14–15 cm) and MBC3 (19–20 cm), where n D 1.

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

9

Fig. 4. Mean sediment fine fraction trace metal profiles for triplicate cores at MBC1, MBC2 and MBC3. All data are means of triplicate cores, except for MBC2 (14–15 cm) and MBC3 (19–20 cm), where n D 1.

10

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

Table 4 Ranges of mean .n D 3/ mud content, porosity, fine fraction TOC, total sediment TOC and fine fraction carbonate

Mean mud content range (%) (RSD) Mean porosity range (%) Mean fine fraction TOC range (%) Mean bulk sediment TOC range (%) Mean fine fraction carbonate range (%)

MBC1

MBC2

MBC3

4.2–13.3 (14%) 37–50 14.0–19.2 3.4–4.7 23.8–26.0

9.0–22.9 (37%) 27–68 15.0–16.9 3.3–5.7 23.8–24.7

16.4–24.3 (19%) 43–62 14.7–16.7 4.5–6.4 24.2–25.3

Pb, Zn and possibly Ni into the dissolved-phase is an alternative explanation. Downcore fine fraction trace metal concentrations at MBC2 and MBC3 show little variation, with only Pb displaying a slight downcore increase below 5 cm. Although bulk sediment EDTA-extractability (ratio of EDTA-extractable metal : total extractable metal) of trace metals is highly variable, there is a similar gradient in the extractable concentrations in bulk sediment and in the fine fraction (Fig. 5). Mean Cu, Pb, Zn and Ag EDTA extractability is highest at MBC1, indicating that these trace metals are possibly more bioavailable at this location than at greater distances from the outfall. There is no correlation between EDTA-extractable trace metals and the mud content in the sediments for all elements analysed (R 2 < 0:18; p > 0:05; n D 50), except Fe which displays a weakly positive correlation (R 2 D 0:37; p < 0:05; n D 50). Bulk sediment EDTA-extractable trace metal analysis is therefore a valid technique in identifying enrichment in sediments adjacent to the outfall and is independent of textural variability. The mean EDTA extractability at each location follows the order Ag D Pb > Cu D Zn > Mn > Fe (Table 5). Co and Ni were below the detection limit in all samples analysed, indicating that the extractability of both of these trace metals is <10% of total sediment concentrations. This supports the view that both Co and Ni are mainly associated with the refractory mineral fraction of the sediments, rather than with the loosely bound fraction. 4.3. Dissolved-phase metals During the initial stages of diagenesis, the degradation of organic matter and the transition from ox-

Table 5 Mean EDTA extractability of five bulk sediment samples (0–10 cm) from each of the triplicate cores at the three sampling sites MBC1, MBC2 and MBC3 (all values are per cent of total extractable trace metals; numbers in brackets indicate SD) Element

MBC1

MBC2

MBC3

Ag Pb Cu Zn Mn Fe

86 (13) 82 (9) 31 (8) 28 (4) 17 (2) 6 (2)

68 (14) 82 (13) 23 (11) 22 (5) 18 (6) 8 (5)

66 (7) 65 (7) 21 (5) 18 (3) 14 (3) 6 (2)

idizing to reducing conditions has a major influence on the partitioning of Fe, Mn and bioreactive trace metals (Widerlund, 1996). In marine sediments, oxic respiration, nitrate reduction, iron and manganese reduction, methane production and sulphate reduction all contribute to organic carbon oxidation. Based on the steepness of downcore porewater Fe and Mn gradients, the reduction of Fe and Mn oxides has been thought to be of minor importance in controlling carbon oxidation in marine environments (Bender and Heggie, 1984; Bender et al., 1989). However, microbiological reduction of Fe and Mn oxides from a wide variety of organic substrates can be a significant process in anaerobic, organic matter-rich marine sediments (Lovley and Phillips, 1988; Myers and Nealson, 1988). Porewater Fe, Mn and Cu concentrations are highly variable both with depth in the sediment, as well as laterally (Fig. 6). Overlying water Fe, Mn and Cu concentrations are lower than the surficial sediment porewater concentrations and are close to detection limits. Downcore dissolved-phase Mn profiles are similar in all three cores, but differ in magnitude and vertical shift. Manganese concentrations in core MBC1A

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

11

Fig. 5. Mean EDTA extractability of bulk sediment trace metals compared to total extractable trace metals at MBC1, MBC2 and MBC3. All data are means of triplicate cores, except for MBC2 (14–15 cm) and MBC3 (19–20 cm), where n D 1.

decrease from 0.36 µmol l 1 at the surface to 0.22 µmol l 1 at 9 to 10 cm. The overlying water Mn concentration in core MBC1A is 0.001 µmol l 1 . Cores MBC2A and MBC3A display a subsurface increase from 0.51 µmol l 1 at the surface to 0.92 µmol l 1 at 1 to 2 cm depth, and from 0.26 µmol l 1 to 1.35 µmol l 1 at 2–3 cm, before decreasing down the cores to 0.07 µmol l 1 and 0.49 µmol l 1 , respectively. The overlying water Mn concentrations in cores MBC2A and MBC3A are 0.024 µmol l 1 and 0.004 µmol l 1 , respectively.

The Mn profiles at the three sites are characteristic of anoxic and suboxic sediments where oxidant consumption during organic matter oxidation, in a thin zone near the sediment–water interface leads to a reduced Mn flux out of the sediments (Froehlich et al., 1979; Aller, 1980; Elderfield et al., 1981; Heggie et al., 1987). The decrease in the dissolved Mn concentration with depth, as observed in all three cores, is common in sedimentary porewaters and results from solid-phase Mn oxide dissolution at depth and a migration of reduced porewater Mn to surface oxic

12

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

Fig. 6. Fe, Mn and Cu porewater profiles and overlying water concentrations of cores MBC1A, MBC2A and MBC3A. The dashed vertical line in the Cu profile graph marks the detection limit of 5 nmol l 1 .

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

sediments. The field data indicate the presence of a redox boundary at less than 1 cm at MBC1, deepening to 1–2 cm and 2–3 cm at MBC2 and MBC3, respectively. This trend indicates deeper oxygenation of the surficial sediments with greater distance from the outfall. The surface enrichment of EDTA-extractable Mn, observed at all three sites, is possibly a result of oxidation and immobilization of dissolved Mn diffusing from deeper sediments (Burdige and Gieskes, 1983). The inferred depth of the redox boundary is consistent with oxygen penetration of only 6 mm into the surficial sediments on the middle shelf adjacent to Sydney (G.P. Bickford, pers. commun., 1996). Dissolved Fe profiles in cores MBC1A and MBC2A are similar; concentrations in the surficial 2 cm are below 0.5 µmol l 1 , increasing sharply to 3.3 µmol l 1 and 2.2 µmol l 1 , respectively, below the redox boundary (2–3 cm), and decreasing gradually to below 1.0 µmol l 1 at 9–10 cm. Core MBC3A has a markedly higher dissolved-phase Fe concentration than the other two cores, with concentrations increasing from 1.2 µmol l 1 to 9.2 µmol l 1 at 1–2 cm; a slight decrease to 4.3 µmol l 1 at 4–5 cm is followed by a sharp increase to greater than 30 µmol l 1 at 9–10 cm, before levelling off to <10 µmol l 1 at 14–15 cm. Cores MBC1A and MBC2A show no such vertical progression of Fe mobilization at depth, whereas core MBC3A displays an increase in dissolved-phase Fe concentrations to a depth of approximately 10 cm. The overlying water Fe concentrations in cores MBC1A, MBC2A and MBC3A are 0.03 µmol l 1 , 1.12 µmol l 1 and 1.77 µmol l 1 , respectively. Manganese reduction precedes Fe reduction in the oxidation sequence of organic carbon (Froehlich et al., 1979). The observed porewater profiles indicate that at MBC1, MBC2 and MBC3 the zone of Fe reduction is approximately 2 cm, 2 cm and 5–8 cm deep, respectively, which is deeper than the zone of Mn reduction. Hence, the theoretical suboxic organic matter oxidation sequence of Mn and Fe mobilization probably applies at all three locations (Froehlich et al., 1979; Alongi et al., 1996; Reimers et al., 1996). Copper concentrations in the porewaters are below 43 nmol l 1 in all samples. The concentrations

13

in core MBC1A decrease from 42 nmol l 1 at the surface to below detection (5 nmol l 1 ) at 9 to 10 cm. In core MBC2A, dissolved-phase Cu also decreases from 13–16 nmol l 1 at the surface to 6 nmol l 1 at 9 to 10 cm depth. The dissolvedphase Cu concentrations in core MBC3A are below detection at the surface (0 to 1 cm depth) and exhibit two peaks at 2 to 3 cm (18 nmol l 1 ) and 9 to 10 cm (33 nmol l 1 ) depth. The overlying water Cu concentrations in cores MBC1A, MBC2A and MBC3A are 9 nmol l 1 , <5 nmol l 1 and 5 nmol l 1 , respectively. Dissolved-phase Cu at MBC1 is enriched in the oxic zone (0–1 cm) owing to strong Cu mineralization (Klinkhammer et al., 1982; Schwedhelm et al., 1988; Petersen et al., 1995). In the deeper anoxic layers (1–10 cm), the degradation of organic matter is limited by the supply of less oxidizing agents, leading to the preferential association of Cu with solid phases. Cu solubility decreases across the redox interface, indicating that this element is not involved in redox cycling, but tends to be enriched in the sediments (Jacobs et al., 1985); this is supported by a pronounced increase in the mean fine-fraction Cu concentration in the surficial sediments at MBC1 and MBC2, relative to MBC3. Widerlund (1996) has shown that Cu displays a nutrient-like behaviour in the marine environment and the early diagenetic remobilization of organicassociated Cu into the dissolved phase exerts a strong control on the Cu flux across the sediment–water interface (Heggie, 1983; Westerlund et al., 1986; Heggie et al., 1987; Shaw et al., 1990). The contribution of reactive, sewage-derived organic carbon to the surficial sediments therefore increases the remobilization potential of bioreactive Cu at MBC1 (Paulson et al., 1991). Dissolved-phase Cu profiles in cores MBC2A and MBC3A indicate the presence of a shallow redox boundary in these sediments at 2–3 cm depth. An elevated dissolved-phase Cu concentration at 9–10 cm depth in core MBC3A coincides with an elevated dissolved-phase Fe concentration (Cr has also shown to be elevated at this depth; C. Matthai, unpubl. data), indicating a possible diagenetic relationship between these elements.

14

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

4.4. Dissolved-phase fluxes

Table 7 Calculated benthic fluxes at the three sites

Steady-state benthic flux estimates are calculated using Fick’s First Law of Diffusion:

Core

F jsed (Fe) (µmol m

dC j F j D D sed (2) j dx where F j is the flux of metal, which is calculated from the product of the effective sediment diffusion coefficient .D sed j / and the concentration gradient .dC j =dx/. The assumptions inherent in this simplified equation have been discussed in detail by a number of authors (Manheim, 1970; Sweerts et al., 1991; Petersen et al., 1995). The concentration gradient is calculated from changes in the porewater trace metal concentration in the uppermost centimetre of the sediment. It is estimated as twice the difference between the porewater concentration in the 0–1 cm sampling interval and the overlying water, therefore assuming a constant dissolved-phase concentration in the upper centimetre (Klinkhammer, 1980). was estimated, after Petersen In this study, D sed j et al. (1995), by correcting the self-diffusion coefficients for divalent cations in water at 18ºC for effects of temperature and porosity, and calculating Dsed j according to the empirical relationship of Sweerts et al. (1991):

MBC1A MBC2A MBC3A

D sed j D D0 . 0:73 C 2:17/

5. Conclusions

1

(3)

where  is sediment porosity and D0 is the temperature-corrected self-diffusion coefficient after Li and Gregory (1974). The temperature- and porosity-corrected Dsed j , as determined from Eq. 3, yield different values for each of the three elements and three locations (Table 6). Surficial sediment porosity varies from 50% at MBC1, to 65% at MBC2 and 52% at MBC3. Table 6 Effective sediment diffusion coefficients for Cu, Fe and Mn at the three sample sites (D0 after Li and Gregory, 1974) Core MBC1A MBC2A MBC3A

D sed j (Fe) (10 6 cm2 s 1 )

D sed j (Mn) (10 6 cm2 s

1.61 2.25 1.70

1.60 2.23 1.68

1)

D sed j (Cu) (10 6 cm2 s 1 ) 1.63 2.28 1.71

9.8 4.7 9.5

2

d

1)

F jsed (Mn) F jsed (Cu) 2 1 (µmol m d ) (µmol m 2 d 1 ) 9.8 19.3 7.4

1065 446 4

The resulting steady-state benthic fluxes, calculated according to Eq. 2, are tabulated in Table 7. The calculated dissolved-phase fluxes of Fe, Mn and Cu are in the low µmol m 2 d 1 ranges, with Fe at MBC2 displaying a negative flux, and the Cu flux at MBC3 being close to zero. The low dissolved-phase benthic fluxes indicate that the surficial sediment at all three locations acts as an efficient trap for dissolved-phase metals. Most organic matter-bound trace metals released during oxic and suboxic organic matter diagenesis are probably scavenged during suboxic diagenetic sulphide mineralization; however, the small dissolved-phase benthic Cu fluxes at MBC1 and MBC2, compared to MBC3, are likely to be linked to decomposition of sewage organic matter and incomplete sequestering of Cu by the solid phases.

Total sediment trace metal concentrations in surficial sediments at MBC1, MBC2 and MBC3 are very low and are mainly controlled by textural changes. Small-scale spatial variability in the mud content and total sediment trace metal concentrations is high at MBC2 because this site is at the boundary of two lithofacies. Sediments near the Malabar deepwater ocean outfall are enriched in trace metals in the fine fraction (<63 µm). The downcore increase in fine fraction trace metal concentrations not only suggests the presence of a deep physical mixing zone, extending beyond at least 10 cm, but also that there may be surficial remobilization of solid-phase trace metals into the dissolved-phase. As a percentage of total extractable total sediment concentrations, weak acid leachable Ag, Cu, Pb and Zn concentrations are higher at MBC1 than at MBC2

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

and MBC3, suggesting that the sewage outfall may be a source of trace metals. Fe and Mn porewater profiles reveal the presence of a subsurface redox boundary at progressively greater depths in the sediment with increasing distance from the outfall. The vertical profiles of porewater Fe and Mn at the three sites are consistent with the zonation model of successive mineralization processes predicted to occur in suboxic sediments (Froehlich et al., 1979; Shaw et al., 1990; Alongi et al., 1996). The presence of elevated dissolved-phase Fe and Mn at depths greater than 1 to 2 cm below the sediment–water interface emphasises that early diagenetic suboxic to anoxic organic matter reduction is an important, if not the dominant, diagenetic process in surficial sediments on the central NSW continental shelf. The very low dissolved-phase Fe and Mn concentrations in porewaters in the surficial 1–2 cm, on the other hand, indicate that near the sediment– water interface there is a thin oxic layer in which oxic respiration is likely to control organic matter decomposition. Decreasing dissolved-phase Cu solubility across the redox boundary indicates that Cu, unlike Fe and Mn, is not involved in redox cycling. Dissolved-phase fluxes of Fe, Mn and Cu from the early diagenetic remobilization of organic matter are very low at all three sites, suggesting that most of the dissolved-phase metals are sequestered by lowsolubility sulphide minerals before they can escape from the sediments. If this is correct, these sediments are an efficient sink for trace metals.

Acknowledgements Analyses of dissolved-phase trace metals were undertaken at the ANSTO laboratory (Environment Division) under the supervision of R. Szymczak.

References Aller, R.C., 1980. Diagenetic processes near the sediment–water interface of Long Is. Sound, II. Fe and Mn. Adv. Geophys. 22, 351–415. Alongi, D.M., Boyle, S.G., Tirendi, F., Payn, C., 1996. Compo-

15

sition and behaviour of trace metals in post-oxic sediments of the Gulf of Papua, Papua New Guinea. Estuarine Coastal Shelf Sci. 42, 197–211. Balls, P.W., 1989. The partition of trace metals between dissolved and particulate phases in European coastal waters: a compilation of field data and comparison with laboratory studies. Neth. J. Sea Res. 23, 7–14. Bender, M.L., Heggie, D.T., 1984. Fate of organic carbon reaching the deep sea floor: a status report. Geochim. Cosmochim. Acta 48, 977–986. Bender, M.L., Jahnke, R., Weiss, R., Martin, W., Heggie, D.T., Orchardo, J., Sowers, T., 1989. Organic carbon oxidation and benthic nitrogen and silica dynamics in San Clemente Basin, a continental borderland site. Geochim. Cosmochim. Acta 53, 685–697. Birch, G.F., 1981. The Karbonat-Bombe: a precise, rapid and cheap instrument for determining calcium carbonate in sediments and rocks. Trans. Geol. Soc. S. Afr. 84, 199–203. Birch, G.F., 1996. Sediment-bound metallic contaminants in Sydney’s estuaries and adjacent offshore, Australia. Estuarine Coastal Shelf Sci. 42, 31–44. Birch, G.F., Davey, S., 1995. Accumulation of metallic contaminants in surficial sediments on a high energy continental shelf (Sydney, Australia). Sci. Total Environ. 170, 81–93. Bryan, G.W., 1984. Pollution due to heavy metals and their compounds. In: Kinne, O. (Ed.), Marine Ecology. Wiley, Chichester, Vol. 5, pp. 1289–1431. Burdige, D.J., Gieskes, J.M., 1983. A porewater=solid-phase diagenetic model for manganese in marine sediments. Am. J. Sci. 283, 29–47. Calmano, W., Wolffgang, A., Fo¨rstner, U., 1990. Exchange of heavy metals between sediment components and water. In: Broekaert, J.A.C., Gucer, S., and Adams, F. (Eds.) Metal Speciation in the Environment. NATO ASI Ser. G23, SpringerVerlag, Berlin, pp. 503–522. Canfield, D.E., 1989. Sulfate reduction and oxic respiration in marine sediments: implications for organic carbon preservation in euxinic environments. Deep Sea Res. 36, 121–138. Chapman, P.M., Paine, M.D., Arthur, A.D., Taylor, L.A., 1996. A triad study of sediment quality associated with a major, relatively untreated marine sewage discharge. Mar. Pollut. Bull. 32, 47–64. Dankers, N., Laane, R., 1983. A comparison of wet oxidation and loss on ignition of organic material in suspended matter. Environ. Technol. Lett. 4, 283–290. Davies, P.J., 1979. Marine geology of the continental shelf off southeast Australia. Bur. Miner. Resour., Geol. Geophys. Bull. 195, 51 pp. di Toro, D.M., Mahony, J.D., Hansen, D.J., Scott, K.J., Carlson, A.R., Ankley, G.T., 1992. Acid volatile sulfide predicts the acute toxicity of cadmium and nickel in sediments. Environ. Sci. Technol. 26, 96–101. Elderfield, H., McCaffrey, R.J., Luedtke, N., Bender, M., Truesdale, V.W., 1981. Chemical diagenesis in Narragansett Bay sediments. Am. J. Sci. 281, 1021–1055. EPA, 1995. Sydney Deepwater Outfalls Final Report Series Vol-

16

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17

ume 2: Sewage Plume Behaviour. NSW Environment Protection Authority Report No. EPA 95=41. EPA, 1996. Sydney Deepwater Outfalls Final Report Series Volume 1: Assessment of the Deepwater Outfalls. NSW Environment Protection Authority Report No. EPA 96=17, March 1996, 194 pp. Fagan, P., Misciewicz, A.G., Tate, P.M., 1992. An approach to monitoring sewage outfalls — a case study on the Sydney deepwater outfalls. Mar. Pollut. Bull. 25, 172–180. Froehlich, P.N., Klinkhammer, G.P., Bender, M.L., Luedtke, N.A., Heath, G.R., Culler, D., Dauphin, P., Hammond, D., Hartmann, B., Maynard, V., 1979. Early oxidation of organic matter in pelagic sediments of the eastern equatorial Atlantic: suboxic diagenesis. Geochim. Cosmochim. Acta 43, 1075– 1090. Gordon, A.D., Hoffman, J.G., 1986. Sediment features and processes of the Sydney continental shelf. In: Frankel, E., Keene, J.B., Waltho, A.E. (Eds.), Recent Sediments in Eastern Australia: Marine Through Terrestrial. Geological Society of Australia (N.S.W. Division), Sydney, pp. 29–51. Heggie, D.T., 1983. Copper in the Resurrection Fjord, Alaska. Estuarine Coastal Shelf Sci. 17, 613–635. Heggie, D.T., Klinkhammer, G.P., Cullen, D., 1987. Manganese and copper fluxes from continental margin sediments. Geochim. Cosmochim. Acta 51, 1059–1070. Irvine, I.A., 1980. Sydney Harbour: Sediments and Heavy Metal Pollution. Ph.D. thesis (unpubl.), University of Sydney. Jacobs, L., Emerson, S., Skei, J., 1985. Partitioning and transport of metals across the O2 =H2 S interface in a permanently anoxic basin: Framvaren Fjord, Norway. Geochim. Cosmochim. Acta 49, 1433–1444. Kern, U., Westrich, B., 1995. Sediment contamination by heavy metals in a lock-regulated section of the River Neckar. Mar. Freshwater Res. 46, 101–106. Klinkhammer, G.P., 1980. Early diagenesis in sediments from the eastern equatorial pacific, II. Pore water metal results. Earth Planet. Sci. Lett. 49, 81–101. Klinkhammer, G.P., Heggie, D.T., Graham, D.W., 1982. Metal diagenesis in oxic marine sediments. Earth Planet. Sci. Lett. 61, 211–219. Li, Y.H., Gregory, S., 1974. Diffusion of ions in seawater and in deep sea sediments. Geochim. Cosmochim. Acta 38, 703–714. Lovley, D.R., Phillips, E.J.P., 1988. Novel mode of microbial metabolism: organic carbon oxidation coupled to dissimilatory reduction of iron and manganese. Appl. Environ. Microbiol. 54, 1472–1480. Manheim, F.T., 1970. The diffusion of ions in unconsolidated sediments. Earth Planet. Sci. Lett. 9, 307–309. Matthai, C., Birch, G.F., 1995. Natural and anthropogenic heavy metal distribution in surficial sediments on the central NSW continental shelf. In: 29th Newcastle Symposium on Advances in the Study of the Sydney Basin. Department of Geology, The University of Newcastle, pp. 74–81. Matthai, C., Birch, G.F., 1996. Trace metals in surficial marine sediments near a large ocean outfall off Malabar, Sydney, Australia. In: Continental Shelves in the Quaternary: Inter-

pretation, Correlation, Application, IGCP 396. Department of Geology and Geophysics, The University of Sydney, 70 pp. Myers, C.R., Nealson, K.H., 1988. Microbial reduction of manganese oxides: interactions with iron and sulfur. Geochim. Cosmochim. Acta 52, 2727–2732. O’Reilly Wiese, S.B., MacLeod, C.L., Lester, J.N., 1997. Partitioning of metals between dissolved and particulate phases in the salt marshes of Essex and North Norfolk (UK). Environ. Technol. Lett. 18, 399–408. Paulson, A.J., Curl Jr., H.C., Cokelet, E.D., 1991. Remobilization of Cu from marine particulate organic matter and from sewage. Mar. Chem. 33, 41–60. Petersen, W., Wallmann, K., Pinglin, L., Schroeder, F., Knauth, H.-D., 1995. Exchange of trace elements at the sediment– water interface during early diagenesis processes. Mar. Freshwater Res. 46, 19–26. Ravizza, G.E., Bothner, M.H., 1996. Osmium isotopes and silver as tracers of anthropogenic metals in sediments from Massachusetts and Cape Cod bays. Geochim. Cosmochim. Acta 60, 2753–2763. Reimers, C.E., Ruttenberg, K.C., Canfield, D.E., Christiansen, M.B., Martin, J.B., 1996. Porewater pH and authigenic phases formed in the uppermost sediments of the Santa Barbara Basin. Geochim. Cosmochim. Acta 60, 4037–4057. Roy, P.S., 1980. Regional geology of the central and northern New South Wales coast. Geol. Jahrb. 56, 25–36. Roy, P.S., 1985. Marine sand bodies on the south Sydney shelf, S.E. Australia. Coastal Studies Unit, University of Sydney, Report 85=1. Scanes, P.R., Philip, N., 1995. Environmental impact of deepwater discharge of sewage off Sydney, NSW, Australia. Mar. Pollut. Bull. 31, 343–346. Schneider, P.M., Davey, S.B., 1995. Sediment contaminants off the coast of Sydney, Australia: a model for their distribution. Mar. Pollut. Bull. 31, 262–272. Schults, D.W., Ferraro, S.P., Smith, L.M., Roberts, F.A., Poindexter, C.K., 1992. A comparison of methods for collecting interstitial water for trace organic compounds and metal analyses. Water Res. 26, 989–995. Schwedhelm, E., Vollmer, M., Kersten, M., 1988. Bestimmung von Konzentrationsgradienten gelo¨ster Schwermetalle an der Sediment=Wasser-Grenzfla¨che mit Hilfe der Dialysetechnik. Fresenius Z. Anal. Chem. 332, 756–763. Shaw, T.J., Gieskes, J.M., Jahnke, R.A., 1990. Early diagenesis in differing depositional environments: the response of transition metals in pore water. Geochim. Cosmochim. Acta 54, 1233– 1246. Skowronek, F., Sagemann, J., Stenzel, F., Schulz, H.D., 1994. Evolution of heavy-metal profiles in River Weser estuary sediments, Germany. Environ. Geol. 24, 223–232. Sweerts, J.P.R.A., Baer-Gilissen, M.J., Cornelese, A.A., 1991. Oxygen-consuming processes at the profundal and littoral sediment–water interface of a small meso-eutrophic lake (Lake Vechten, The Netherlands). Limnol. Oceanogr. 36, 1124–1133. Sydney Water, 1996. Ecological and human health risk assessment of chemicals in sewage discharges to ocean waters. Sydney Water Corp., January 1996.

C. Matthai et al. / Journal of Geochemical Exploration 64 (1998) 1–17 Webster, J., Ridgway, I., 1994. The application of the equilibrium partitioning approach for establishing sediment quality criteria at two UK sea disposal and outfall sites. Mar. Pollut. Bull. 28, 653–661. Westerlund, S.F.G., Anderson, L.G., Hall, P.O.G., Iverfeldt, A., Van der Loeft, M.M.R., Sundby, B., 1986. Benthic fluxes of cadmium, copper, nickel, zinc and lead in the coastal

17

environment. Geochim. Cosmochim. Acta 50, 1289–1296. Widerlund, A., 1996. Early diagenetic remobilization of copper in near-shore marine sediments: a quantitative pore-water model. Mar. Chem. 54, 41–53. Ying, W., Batley, G.E., Ahsanullah, M., 1992. The ability of sediment extractants to measure the bioavailability of metals to three marine invertebrates. Sci. Total Environ. 125, 67–84.