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Behaviour of heavy metals in a mud flat of the Scheldt estuary, Belgium
Marine Pollution Bulletin Marine Pollution Bulletin, Volume 22, No. 3, pp. 128-134. 1991. Printed in Great Britain.
0025-326X/91 $3.01)+0.00 © 1991 Pergamon Press plc
Behaviour of Heavy Metals in a Mud Flat of the Scheldt Estuary, Belgium S. PANUTRAKUL* and W. BAEYENS
Laboratory of Analytical Chemistry and Geochemistry, University of Brussels (VUB), Pleinlaan 2, B-1050 Brussels, Belgium *Permanent address: Phuket Marine Biological center, P.O. Box 60, Phuket 83000, Thailand
The Ballasplaat intertidal mud flat in the Scheldt estuary has been polluted by Cd and Pb and also high amounts of organic matter as a result of suspended matter deposition. The degree of pollution is, however, not uniform over the mud flat due to varying physicochemical conditions (essentially variable redox profiles). Measurements of the redox profile, and the metal concentrations in the pore water, the total sediment and the fraction < 63 jim combined with sequential extraction results and enrichment factors can explain the behaviour of Cd, Pb, Fe, and Mn in the sediments. In a strong anoxic sediment, where sulphate reducing bacteria are active, the heavy metals are trapped as poorly soluble metal complexes while in an oxic or suboxic sediment metals tend to be redissoived due to the oxidation of organic matter, and the reduction of particulate Fe and Mn oxyhydroxides. The transportation of heavy metals across the sediment water interface is very much depending upon the physicochemical conditions of the sediment.
Early diagenetic processes control the behaviour, and more specifically the dissolved-solid phase distribution of the heavy metals in sediments. The key process is the bacterial degradation of organic matter and the simultaneous consumption of an oxydant. Once oxygen is exhausted, manganese (IV), nitrate, iron (III) and sulphate are successively used as electron acceptors (Alexander, 1977; Froelich et al., 1979; Sundby et al., 1981; Jorgensen, 1989). In this study, the Ballastplaat intertidal mud fiat, which is located at about 60 km from the Western Scheldt river mouth, has been sampled (Fig. 1). The Scheldt originates in Saint-Quentin (France), passes through Belgium and the Netherlands, and flows into the North Sea at Vlissingen. The tidal amplitude varies between 4 m near Vlissingen, 5 m near Antwerp and 2 m near Gent. The river discharge varies from 40 to 350 m 3 s-1 with an average of 90 m 3 s-1 at Antwerp which is rather small; hence the influence of the tide and salt intrusion is very important in this estuary. 128
Baeyens et al. (1988) subdivided the Scheldt estuary into three different zones based on their physical and chemical characteristics. The Western Scheldt has been reported as a heavily polluted estuary since it receives domestic and industrial waste water from cities such as Brussels, Antwerp and Gent. The low water quality is very stressed in the upstream estuary where the sampling site is located (Somville & De Pauw, 1982). The Ballasplaat mud flat is separated into two parts by a dike (Fig. 1) which has some effect on the sedimentation pattern. The sediment compositions on both sides of the dike are different; the sandy part (upstream of the dike) and the muddy part (downstream of the dike) are called station A and station B, respectively, in this work. Van Impe (1985) demonstrated in his work that the species diversity of macrofauna on this mud flat and surrounding areas has receded. P. Meire (unpublished data) showed that the situation of the benthic communities between the two stations A and B is not the same; station A is slightly better than station B. The aim of our work was to study the heavy metals concentrations on the mud flat in relation to the redox potential, the organic matter and A1 content (often used as parameters for the fine sediment fraction) and the sequential analyses results. Materials and M e t h o d s Sediment cores were collected in February 1989. The Plexiglas tubes 30 cm in length and 4 cm in diameter were pushed into the sediments, stoppered at the top while avoiding air between the overlying water and the caps, carefully pulled out of the sediments and immediately stoppered at the bottom. Although these cores were tightly capped, they were in addition wrapped and tightly taped with plastic to screen them from contact with the air and then transported to the analytical chemistry laboratory, University of Brussels (VUB), immediately after sampling. The cores were subsampled and centrifugated in a glove box containing nitrogen atmosphere. In only a few subsamples at station A interstitial water could be separated. The pore water samples were immediately
Volume 22/Number 3/March 1991 TABLE 1
The sequential extraction procedure used in this experiment. Step
Extractant
Fraction
1
The exchangeableand carbonate fraction
2
The reducible fraction
3 4
The acid solublefraction The oxidizablefraction
5
The residual fraction
40 ml of I M Sodium acetate buffered to pH 5 with acetic acid 40 ml of 1 M Hydroxylaminebuffered to pH 5 with citric acid 33 ml of 0.01 M HNO3 100 ml of 1 M Ammoniumacetate in 6% HNO3 Aqua regia:HF ratio 8 :2
Process Shaking twice, 2 h each Shaking twice, 2 h each Shaking 3 times, 1 h each Digesting twice with 100 ml of 30% H2O 2 buffered to pH 2 with HNO3.Then shaking the residue with the extractant overnight. Acid digestion procedure as described in Materials and .Methods
1. Centrifugation was carried out after each extraction to separate the solid and the aqueous phase. The solutions were used to determine the metals AI, Fe and Mn by F-AAS, and Cd, Pb by ET-AAS.The residue from each step was used for the next step. 2. The residues from step 4 were dried by lyophilization before digestion in step 5. 3. Calculationis based on dry wt.
Fig. 1 The sampling sites on the Ballasplaat mud flat in the Scheldt
Estuary; stations A and B are separated by a dike. analysed for dissolved Cd and Pb. The sediment subsamples were deep-frozen in polyethylene (EE.) bottles and dried by lyophilization. A n aliquot of each dried sample was ground and homogenized with an agate m o r t a r while another fraction was sieved through a nylon sieve (63 ~tm mesh size) to isolate the < 63 ~tm fraction. The sediment samples were mineralized with an acid digestion for the analysis of Fe, Mn, Cd, and Pb. A b o u t 0.25 g of homogenized samples was weighed in a hot acid cleaned Teflon b o m b (PTFE). 8 and 2 ml of aqua regia and concentrated suprapur H F were respectively added. These b o m b s were tightly closed and the digestion was carried out overnight at 60°C. Then the samples were cooled in a deep-freeze and slowly
evaporated on a hot plate. T h e residue was diluted and brought to 25 ml with 4% HNO3. These solutions were kept in acid-cleaned polyethylene bottles prior to analysis. A1 was determined after an alkaline fusion of the sediment. About 0.1 g solid sample was mixed with 0.25 g of lithium metaborate (LiBO2) in a platinum crucible; then the mixture was fused at 1100°C for 1 h. The fused material was immediately cooled to room temperature, then dissolved with 4% H N O 3 on a hot plate (at 100°C) with a magnetic stirrer and brought to 100 ml in a volumetric flask. The solutions were stored in acid-cleaned bottles before the measurement of AI. AI, Fe, and Mn were determined by flame atomic absorption spectrophotometry (F-AAS) while Cd and Pb were determined by electrothermal atomic absorption spectrophotometry (ET-AAS) with Z e e m a n correction. Reference materials (SL 1 and SL 5 from I.A.E.A. in Wien, NBS 1646 from NBS in Washington, MESS 1 from the National Research Council Canada in Ottawa and B C R 143 from the C E C in Brussels) were regularly cycled through the whole analytical procedure in order to validate the results. The recovery for the various metals yield the following results: A1 (%) SL 5: 8.19 + 0.28, found 8.33 + 0.49; NBS 1646: 6 . 2 5 + 0 . 2 0 , found 6 . 2 4 + 0 . 0 8 ; MESS 1: 5 . 8 3 + 0 . 2 0 , found 5.93_+0.1, Fe (%) SL h 6.74_+0.17, found 6.39; SL 5: 4.45+_0.19, found 4.58-+0.24; NBS 1646: 3.35 -+ 0.10, found 3.50; MESS 1:3.049 -+ 0.179, found 2 . 7 3 + 0 . 2 , Mn (rag kg -1) SL 1: 3 4 0 + 1 6 , found 349_+3; SL 5: 8 5 2 + 3 7 , found 8 5 3 + 6 1 ; NBS 1646: 375-+20, found 357; MESS 1: 510_+30, found 459-+14, Cd (mg kg - l ) MESS 1: 0.59_+0.9, found 0.62 -+ 0.12; B C R 143:31.1 _+ 1.2, found 30.5 -+ 0.9, Pb (mg kg -l) SL 1: 37.7_+7.4, found 36.5_+3; SL 5: 129-+ 26, found 151 _+25; NBS 1646:28.2 _+ 1.8, found 27.0; MESS 1:34.0 _+6.1, found 32.3 _+4.8. The agreement between the certified and our values was generally good, except for Fe and Mn in the reference material MESS 1, where our values were somewhat lower (around 90%). Cd and Pb in the interstitial water were determined by ET-AAS. All the vessels and materials used for metal determinations were cleaned with acid and dried in a laminar flow hood before use. The ignition loss at 550°C during 4 h was determined as an 129
Marine Pollution Bulletin
indicator of the organic matter content since it has been shown that there is a good correlation between % of ignition loss and organic matter content (Dean, 1974; Byer et al., 1978; Baeyens et al., 1988). In order to get an idea of the metal speciation in a sediment, a sequential extraction method was used. The scheme, shown in Table 1 is a modification of the scheme suggested by Salomons & Forstner (1980). One must be aware, however, that sequential extraction is only a tool to better understand the distribution of an element over the different major sediment fractions as defined in Table 1. BCR of the EC is preparing reference materials in order to express sequential extraction results relative to these standards. Meanwhile, we have tested in detail the reproducibility of the various extraction steps included in our method, as well as the efficiency of each extraction step. These results, which will be published in a related paper (part of them have been presented by Baeyens et al., 1988), allow us to consider the sequential extraction technique as a valuable, additional tool for the interpretation of the geochemical behaviour of metals in sediments.
Results and Discussion The mud fiat has a heterogeneous composition showing fine sediment at station B and coarser sediment at station A. Sediment A has a thin dark brown surface layer composed of fine material while below coarse brown material is present. Sediment B consists of fine, dark brown material except for a black layer at 3-5 cm depth with a strong smell of hydrogen sulphide. 93.5% of sediment A and 39.8% of sediment B have a granulometric size < 63 ~tm. In sediment B, the redox profile (D. Van Gansbeke, personal communication) steeply drops from the first cm depth and tends to stabilize at --210 mV from the depth of 6 cm on, indicating a shallow switch from an oxic to an anoxic condition (Table 2). In contrast the oxic/anoxic interface is situated at about 8-10 cm depth at station A. The ignition loss profiles of the bulk sediments are shown in Fig. 2; they vary from 1.36 to 3.52% in sediment A and from 4.30 to 10.37% in B. The metal contents at stations A and B in the bulk sediment and in the fraction < 63 ~tm are presented in Figs 2 and 3, respectively. They are qualitatively and quantitatively different between both stations, but at a given station all metals show a similar trend. In sediment A, the metal content increases with increasing depth whereas maximum metal contents at about 3-5 cm depth are observed in sediment B. Cd and Pb showed high levels, 41 gg g-i and 200 ~tg g-i respectively, but Baeyens et al., (1982) reported 19 ~tg g-~ Cd and 174 ~tg g-~ Pb in surface sediments collected in the navigation channel next to the Ballasplaat, and Wollast et al. (1985) reported values of 15 ~tg g-1 Cd and 455 ~tg g-~ Pb in bottom sediments between 50 and 80 km from the fiver mouth of the Scheldt estuary. Obviously, the areas where high metal contents in the Scheldt are reported are situated mainly upstream from the border between Belgium and the Netherlands (Baeyens et al., 1982; Wollast et al., 1985; this work). Downstream of 130
TABLE 2 Dissolved Cd, Pb, and redox potential in the sediments at stations A and B.
Depth
(cm)
Cd (ggl -t)
Station A Pb 0tgl -l)
Eh* (mV)
Cd (~tgl-')
Station Pb (~tgl -~)
2.4 3.4 3.2
+330 +300 +268 +254 +241 +238 +72
2.07 ND ND ND ND ND ND
2.0 1.1 1.0 3.0 2.0 2.2 2.0
B
Eh* (mV)
Surface
water 0-1 1-2 2-3 3-4 4-5 5-7.5 7.5-10 10-15 15-20 20-25
11.25 ND
ND
-
-
- - 5 8
--190 --242
-
ND ND ND
-
+311 --68 --139 --181 --205 --209 --210 --215
5.9 2.8 5.1
ND = Not detectable, - = No Data. *D. Van Gansbeke, personal communication.
this zone, the contamination in the bottom sediments is much less (Araujo et al., 1988). The sequential extraction results at both stations are plotted in Fig. 4. It appears from these results that for Mn the carbonate and exchangeable fraction is a very important one (about 30% in sediment A, 40% in sediment B). None of the other studied metals shows such a high labile sediment fraction. The reducible fraction is relatively high in the oxic layer of sediment A, but decreases rapidly in the deeper anoxic layers. In addition enrichment factors indicate a depletion of Mn in the suboxic sediment (oxygen is here exhausted but sulphate not yet reduced) and enrichment of Mn in the anoxic sediment (this is the sulphate reduction zone). Fe does not show the same behaviour as Mn, since it is mainly associated with the residual fraction (7080%). 3-8% and 5-12% of Fe was found in the exchangeable and carbonate fraction at station A and B, respectively; Fe in this fraction tends to increase with increasing depth. The reducible fraction of Fe is higher in the surface layer (the oxidized Fe species are found in the more oxic surface layers) and lower in the deeper sediment. The acid soluble Fe is higher at all depths in sediment B than A. At both stations low amounts of Fe are found in oxidizable fraction (2-3%). According to the low amount of Fe in the oxidizable fraction, pyrite is not an important fraction of Fe in this mud flat, not even in the anoxic sediment layers. This is in contrast with Lord & Church (1983) who found a large accumulation of pyrite at the O2/H2S interface (above this zone little pyrite accumulated due to extensive reoxidation). The acid soluble Fe is more important than the pyrite fraction in our sediments A and B. In the surface and suboxic sediment layers of sediments A and B, a substantial fraction of the Cd content was found to be associated with the exchangeable and carbonate fraction and the reducible fraction: in sediment A 40-45%, in sediment B about 15%. However, most of the Cd content at both stations was found in the oxidizable fraction, which is theoretically composed of organic-metal and metal-sulphide complexes, showing an increase with increasing depth and decreasing redox potential. Jacobs & Emerson (1982) concluded in their
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Depth (cm) Fig. 2 Metals and organic contents in bulk sediment plotted in function of depth. The profiles show lower metal contents at station A than B; - - W - - station A, - - - 0 - - station B.
work that heavy metals such as Cu, Zn, and Cd display a dramatic solubility decrease across the oxic/anoxic
interface leading to enrichment in the anoxic sediment layers. In our study, no dissolved Cd was observed in the pore waters. This might be a consequence of a very fast formation of a strong cadmium sulphide complex (Emerson et al., 1983; Boulegue, 1983). The enrichment factor for Cd is about 10 times higher in sediment B than A, and is most pronounced at a depth of 5-7.5 131
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132
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Volume 22/Number 3/March 1991
cm in sediment B, where the black colour and strong smell of H2S is observed. The work of Jacobs & Emerson (1982) and Emerson et al. (1983), who studied the partitioning of metals across the O2/H2S interface in a permanently anoxic basin, supports our conclusion that metal sulphide precipitation influences the magnitude of the metal enrichment in sediments. Nevertheless, dissolved Cd was found about five times higher in the overlying water at station A, compared to station B, i.e. 11 ~tg 1-1 and 2 ~tg 1-1 respectively. This is in contrast to the Cd content in the sediments at each station as shown above, but two possible explanations exist. Station A is situated close to the discharge pipe of the BASF factory (producing fertilizers, plastics etc.) and to the discharge point of a small river, the Schijn, which is pumped into the Scheldt. The Schijn flows through the industrial area of Antwerp situated on the right bank of the river. These two outflows may have a strong influence on the local high dissolved Cd values. In addition, sampling was carried out at shallow water depth. Oxidation of reduced cadmium during low tide, may also influence locally the water column concentrations. In contrast to Cd, Pb is mainly associated with the residual fraction (about 60% at both stations). The other 40% is shared by the other fractions, with the acid soluble fraction the smallest. The exchangeable and carbonate fraction and the reducible fraction are more important in the surface layers while in the deeper sediment layers these two fractions diminish in favour of the oxidizable fraction. The size of the residual fraction is hardly affected by the redox conditions. The pore water results at station B show an increase of dissolved Pb with increasing depth. This would indicate that there is a competition between lead sulphide formation and complexation by dissolved organic ligands (e.g. humic acids and others) which control the dissolved/particulate distribution of lead. The metal contents in the fraction < 63 ~tm at station A are higher than in the bulk sediment, but at station B they both show a similar profile. Below 10 cm depth, the metal content in the fine sediment fraction is even higher at station A than at station B. Comparison of metal contents in the sediment fraction < 63 ~tm and in total suspended matter of the Scheldt (Table 3) indicates that the contamination of heavy metals in the fraction <63 p.m is comparable to the suspended matter at the same location, about 60 km from the river mouth. Metal accumulation at the mud flat was created by settling of this contaminated suspended matter. Therefore if the physicochemical conditions at the two
stations were the same, the same range of metal contents in the sediment size fraction < 63 ~m at both stations should be found. This was, however, not the case due to different physicochemical conditions.
Conclusion The contamination at the sampling sites is created by the settling of contaminated suspended matter and by local discharges. A dike, which separates the mud flat in two parts, creates different hydrodynamic conditions and hence different sediment structures and different physicochemical conditions in the sediments at each side of the dike. Station A is located in the sand and coarser grain size area, station B in the mud and finer grain size area. The upper zone of station A is oxic to suboxic while below 1 cm at station B and below 10 cm at station A the sediment is strong anoxic. In the suboxic sediment layers, Fe and Mn were reduced and redissolved but simultaneously other metals such as Pb were also resolubilized. Therefore we found less contamination in the first 10 cm at station A. At station B, which is a favourable place for deposition of fine suspended matter, a strong anoxic condition was emphasized by the imbalance between supply and demand of oxygen. It acts as a trap for metals in the Scheldt estuary due to the formation of poorly soluble metal complexes. The higher metals contamination at this station is very well supported by the observation of lower species diversity and biomass. Finally, we found that the geochemical behaviour of each of the studied metals appears to be different, leading to different sediment profiles in the pore water, the bulk sediment and the sediment fractions. The authors thank the Belgium Ministry for Development and Cooperation (ABOS) and the VUB (FAME) for a research grant for S.P. They are also grateful to F. Dehairs for his comments on the manuscript.
Alexander, M. (1977). Introduction to Soil Microbiology. John Wiley & Sons, Inc., New York. Araujo, M. F. D., Bernard, P. C. & Van Grieken, R. E. (1988). Heavy metal contamination in sediments from the Belgium Coast and Scheldt estuary. Mar. Pollut. Bull. 19, 269-273. Baeyens, W., Wartel, S., Dehairs, E, Decadt, G., Bogaert, M., Gillain, G. & Dedeurwaerder, H. (1982). The river Scheldt as a transport route for heavy metals in the sea. In Distribution, Transport and Fate of Heavy Metals in the Belgian Coastal Marine Environment (A. Disteche & I. Elskens, eds), pp. 87-108. Science Policy, Brussels. Baeyens, W., Monteny, E, Leermakers, M., Lansens, P. & Vandenhoudt, A. (1988). Study of the degree of contamination of dredged sludge in the Scheldt estuary (in Dutch). Synthesis report. University of Brussels (VUB), Brussels.
TABLE 3
Comparison of the metals concentrations in the sediment fraction < 63 p~m in the anoxic sediment layers at stations A and B, and in suspended matter at 60 km from the river mouth. Distance from the river mouth (km) 60 60 60 Station A Station B
AI (mg g-i)
Fe (mg g-i)
Mn (mg g-i)
Cd (mg g-l)
Pb (mg g-i)
57.3 44.8 45.3
55.4 46.8 39.6 38.9
1475 908 772 937
22.4 29.0 22.0 7.45 15.4
174 248 237 186.5 177.7
Reference Dehairs etal. (1986) Baeyens et al. (1982) Baeyens et al. (1988) This work This work
133
Marine Pollution Bulletin Boulegue, J. (1983). Trace Metals (Fe, Cu, Zn, Cd) in Anoxic Environment. Trace Metals in Sea Water (C. S. Wong et al., eds), pp. 563-577. NATO conference series, Plenum Press, New York and London. Byers, S. C., Mills, E. L. & Stewert, P. L. (1978). A comparison of determining organic carbon in marine sediments, with suggestions for a standard method. Hydrologia 58, 43-47. Dean, W. E., Jr. (1974). Determination of carbonate and organic matter in calcareous sediments and sedimentary rocks by loss on ignition: comparison with other methods. J. Sedim. Petrol. 44,242-248. Dehairs, E & Baeyens, W. (1986). Biogeochemical regulation of stable pollutant transfer in open sea and coastal environments. Final report of contract (ENV-766-B), EEC, Brussels. Emerson, S., Jacobs, L. & Lebo, B. (1983). The behavior of heavy metals in marine anoxic water: Solubilities at the oxygen-hydrogen sulfide interface. In Trace Metals in Sea Water (C. S. Wong et al., eds), pp. 579-608. NATO conference series, Plenum Press, New York and London. Froelich, P. N., et al. (1979). Early oxidation of organic matter in pelagic sediments of the eastern equatorial Atlantic: suboxic diagenesis. Geochim. Cosmochim. Acta 43, 1075-1090. Jacobs, L. & Emerson, S. (1982). Trace metals solubility in an anoxic fjord. Earth Planet. Sci. Lett. 60,237-252. Jorgensen, B. B. (1989). Biogeochemistry of chemoautotrophic
Marine I'olh¢tion Bulletin, Volume 22, No. 3. pp. 134-140, 1991. Printed in Great Britain.
bacteria. In Autotrophic Bacteria (H. G. Schlegel & B. Bowien, eds), pp. 117-146. Science Tech Publishers & Springer-Verlag, Madison. Lord, II1, C. J. & Church, T. M. (1983). The geochemistry of salt marshes: Sedimentary ion diffusion, sulfate reduction and pyritization. Geochim. Cosmochim. Acta 47, 1381-1391. Salomons, W. & Forsmer, U. (1980). Trace metal analysis on polluted sediments. Part II: Evaluation of environmental impact. Environ. Techno. Lett. 1,506-517. Somville, M. & De Pauw, N. (1982). Influence of temperature and river discharge on water quality of the Western Schheldt estuary. Water Res. 16, 1349-1356. Strickland, J. D. H. & Parsons, T. R. (1968). A Practical Handbook of Seawater Analysis. Fisheries Research board of Canada. Bulletin No. 167. Sundby, B., Bouchard, G., Lebel, J. & Silverberg, N. (1981). Rates of organic matter oxidation and carbon transport in early diagenesis of marine sediment. Adv. Org. Geochem. 350-354. Van Impe, J. (1985). Estuarine pollution as probable causes of increase of Estuarine birds. Mar. Pollut. Bull. 16, 271-276. Wollast, R., Devos, G. & Hoenig, M. (1985). Distribution of heavy metals in the sediment of the Scheldt estuary. In Progress in Belgian Oceanographic Research (R. van Grieken & R. Wollast, eds), pp. 147-159.
0025-326X/91 S3.00+0.00 © 1991 Pergamon Press plc
Effect of Three Primary Treatment Sewage Outfalls on Metal Concentrations in the Fish Cheilodactylusfuscus Collected Along the Coast of Sydney, Australia C. McLEAN, A. G. MISKIEWICZ and E. A. ROBERTS Scientific Services and the Environmental Projects Group, Water Board, PO Box A53, Sydney, 2000, N.S. W. Australia
Samples of muscle tissue from red morwong Cheilodactylus fuscus (Pisces: Cheilodactylidae) collected at 24 sites along the coast near Sydney Australia were analysed for total concentrations of mercury, arsenic, selenium, zinc, cadmium, lead, nickel and copper. The sites were equally spaced around each of the three major ocean disposal sewage treatment plants (STP) in Sydney located at North Head, Bondi and Malabar. The mean concentrations of mercury, arsenic and zinc were highest in fish caught off Sydney Harbour and Malabar STP. The proportion of fish in which cadmium was detected was also highest off Sydney Harbour and south of Malabar STP. Average selenium concentrations decreased from north to south with peak concentrations at each STP. Many individual values of copper, lead, and nickel were near or below the detection limits and no notable trends were observed for these metals. Mercury was the only metal whose overall mean concentration exceeded
134
the National Health and Medical Research Council (NHMRC) Maximum Residue Limit (MRL).
Worldwide concern has been expressed about the effect of sewage and industrial effluent on the marine environment and numerous studies have been undertaken to assess this. Higher levels of heavy metals such as mercury, lead, copper, zinc and cadmium, have been recorded in the water, sediments and invertebrates near discharge points than at control sites (Greig & Wenzloff, 1977; Roth & Hornung, 1977; Amiel & Navrot, 1978; Young et al., 1981; Smith et al., 1981; Talbot & Chegwidden, 1982; Vasilikiotis et al., 1983; Ward et al., 1986). However usually only mercury was found at higher levels in fish (Dix & Martin, 1975; Young, 1982; Clark & Topping, 1989).
0025-326X/91 $3.01)+0.00 © 1991 Pergamon Press plc
Behaviour of Heavy Metals in a Mud Flat of the Scheldt Estuary, Belgium S. PANUTRAKUL* and W. BAEYENS
Laboratory of Analytical Chemistry and Geochemistry, University of Brussels (VUB), Pleinlaan 2, B-1050 Brussels, Belgium *Permanent address: Phuket Marine Biological center, P.O. Box 60, Phuket 83000, Thailand
The Ballasplaat intertidal mud flat in the Scheldt estuary has been polluted by Cd and Pb and also high amounts of organic matter as a result of suspended matter deposition. The degree of pollution is, however, not uniform over the mud flat due to varying physicochemical conditions (essentially variable redox profiles). Measurements of the redox profile, and the metal concentrations in the pore water, the total sediment and the fraction < 63 jim combined with sequential extraction results and enrichment factors can explain the behaviour of Cd, Pb, Fe, and Mn in the sediments. In a strong anoxic sediment, where sulphate reducing bacteria are active, the heavy metals are trapped as poorly soluble metal complexes while in an oxic or suboxic sediment metals tend to be redissoived due to the oxidation of organic matter, and the reduction of particulate Fe and Mn oxyhydroxides. The transportation of heavy metals across the sediment water interface is very much depending upon the physicochemical conditions of the sediment.
Early diagenetic processes control the behaviour, and more specifically the dissolved-solid phase distribution of the heavy metals in sediments. The key process is the bacterial degradation of organic matter and the simultaneous consumption of an oxydant. Once oxygen is exhausted, manganese (IV), nitrate, iron (III) and sulphate are successively used as electron acceptors (Alexander, 1977; Froelich et al., 1979; Sundby et al., 1981; Jorgensen, 1989). In this study, the Ballastplaat intertidal mud fiat, which is located at about 60 km from the Western Scheldt river mouth, has been sampled (Fig. 1). The Scheldt originates in Saint-Quentin (France), passes through Belgium and the Netherlands, and flows into the North Sea at Vlissingen. The tidal amplitude varies between 4 m near Vlissingen, 5 m near Antwerp and 2 m near Gent. The river discharge varies from 40 to 350 m 3 s-1 with an average of 90 m 3 s-1 at Antwerp which is rather small; hence the influence of the tide and salt intrusion is very important in this estuary. 128
Baeyens et al. (1988) subdivided the Scheldt estuary into three different zones based on their physical and chemical characteristics. The Western Scheldt has been reported as a heavily polluted estuary since it receives domestic and industrial waste water from cities such as Brussels, Antwerp and Gent. The low water quality is very stressed in the upstream estuary where the sampling site is located (Somville & De Pauw, 1982). The Ballasplaat mud flat is separated into two parts by a dike (Fig. 1) which has some effect on the sedimentation pattern. The sediment compositions on both sides of the dike are different; the sandy part (upstream of the dike) and the muddy part (downstream of the dike) are called station A and station B, respectively, in this work. Van Impe (1985) demonstrated in his work that the species diversity of macrofauna on this mud flat and surrounding areas has receded. P. Meire (unpublished data) showed that the situation of the benthic communities between the two stations A and B is not the same; station A is slightly better than station B. The aim of our work was to study the heavy metals concentrations on the mud flat in relation to the redox potential, the organic matter and A1 content (often used as parameters for the fine sediment fraction) and the sequential analyses results. Materials and M e t h o d s Sediment cores were collected in February 1989. The Plexiglas tubes 30 cm in length and 4 cm in diameter were pushed into the sediments, stoppered at the top while avoiding air between the overlying water and the caps, carefully pulled out of the sediments and immediately stoppered at the bottom. Although these cores were tightly capped, they were in addition wrapped and tightly taped with plastic to screen them from contact with the air and then transported to the analytical chemistry laboratory, University of Brussels (VUB), immediately after sampling. The cores were subsampled and centrifugated in a glove box containing nitrogen atmosphere. In only a few subsamples at station A interstitial water could be separated. The pore water samples were immediately
Volume 22/Number 3/March 1991 TABLE 1
The sequential extraction procedure used in this experiment. Step
Extractant
Fraction
1
The exchangeableand carbonate fraction
2
The reducible fraction
3 4
The acid solublefraction The oxidizablefraction
5
The residual fraction
40 ml of I M Sodium acetate buffered to pH 5 with acetic acid 40 ml of 1 M Hydroxylaminebuffered to pH 5 with citric acid 33 ml of 0.01 M HNO3 100 ml of 1 M Ammoniumacetate in 6% HNO3 Aqua regia:HF ratio 8 :2
Process Shaking twice, 2 h each Shaking twice, 2 h each Shaking 3 times, 1 h each Digesting twice with 100 ml of 30% H2O 2 buffered to pH 2 with HNO3.Then shaking the residue with the extractant overnight. Acid digestion procedure as described in Materials and .Methods
1. Centrifugation was carried out after each extraction to separate the solid and the aqueous phase. The solutions were used to determine the metals AI, Fe and Mn by F-AAS, and Cd, Pb by ET-AAS.The residue from each step was used for the next step. 2. The residues from step 4 were dried by lyophilization before digestion in step 5. 3. Calculationis based on dry wt.
Fig. 1 The sampling sites on the Ballasplaat mud flat in the Scheldt
Estuary; stations A and B are separated by a dike. analysed for dissolved Cd and Pb. The sediment subsamples were deep-frozen in polyethylene (EE.) bottles and dried by lyophilization. A n aliquot of each dried sample was ground and homogenized with an agate m o r t a r while another fraction was sieved through a nylon sieve (63 ~tm mesh size) to isolate the < 63 ~tm fraction. The sediment samples were mineralized with an acid digestion for the analysis of Fe, Mn, Cd, and Pb. A b o u t 0.25 g of homogenized samples was weighed in a hot acid cleaned Teflon b o m b (PTFE). 8 and 2 ml of aqua regia and concentrated suprapur H F were respectively added. These b o m b s were tightly closed and the digestion was carried out overnight at 60°C. Then the samples were cooled in a deep-freeze and slowly
evaporated on a hot plate. T h e residue was diluted and brought to 25 ml with 4% HNO3. These solutions were kept in acid-cleaned polyethylene bottles prior to analysis. A1 was determined after an alkaline fusion of the sediment. About 0.1 g solid sample was mixed with 0.25 g of lithium metaborate (LiBO2) in a platinum crucible; then the mixture was fused at 1100°C for 1 h. The fused material was immediately cooled to room temperature, then dissolved with 4% H N O 3 on a hot plate (at 100°C) with a magnetic stirrer and brought to 100 ml in a volumetric flask. The solutions were stored in acid-cleaned bottles before the measurement of AI. AI, Fe, and Mn were determined by flame atomic absorption spectrophotometry (F-AAS) while Cd and Pb were determined by electrothermal atomic absorption spectrophotometry (ET-AAS) with Z e e m a n correction. Reference materials (SL 1 and SL 5 from I.A.E.A. in Wien, NBS 1646 from NBS in Washington, MESS 1 from the National Research Council Canada in Ottawa and B C R 143 from the C E C in Brussels) were regularly cycled through the whole analytical procedure in order to validate the results. The recovery for the various metals yield the following results: A1 (%) SL 5: 8.19 + 0.28, found 8.33 + 0.49; NBS 1646: 6 . 2 5 + 0 . 2 0 , found 6 . 2 4 + 0 . 0 8 ; MESS 1: 5 . 8 3 + 0 . 2 0 , found 5.93_+0.1, Fe (%) SL h 6.74_+0.17, found 6.39; SL 5: 4.45+_0.19, found 4.58-+0.24; NBS 1646: 3.35 -+ 0.10, found 3.50; MESS 1:3.049 -+ 0.179, found 2 . 7 3 + 0 . 2 , Mn (rag kg -1) SL 1: 3 4 0 + 1 6 , found 349_+3; SL 5: 8 5 2 + 3 7 , found 8 5 3 + 6 1 ; NBS 1646: 375-+20, found 357; MESS 1: 510_+30, found 459-+14, Cd (mg kg - l ) MESS 1: 0.59_+0.9, found 0.62 -+ 0.12; B C R 143:31.1 _+ 1.2, found 30.5 -+ 0.9, Pb (mg kg -l) SL 1: 37.7_+7.4, found 36.5_+3; SL 5: 129-+ 26, found 151 _+25; NBS 1646:28.2 _+ 1.8, found 27.0; MESS 1:34.0 _+6.1, found 32.3 _+4.8. The agreement between the certified and our values was generally good, except for Fe and Mn in the reference material MESS 1, where our values were somewhat lower (around 90%). Cd and Pb in the interstitial water were determined by ET-AAS. All the vessels and materials used for metal determinations were cleaned with acid and dried in a laminar flow hood before use. The ignition loss at 550°C during 4 h was determined as an 129
Marine Pollution Bulletin
indicator of the organic matter content since it has been shown that there is a good correlation between % of ignition loss and organic matter content (Dean, 1974; Byer et al., 1978; Baeyens et al., 1988). In order to get an idea of the metal speciation in a sediment, a sequential extraction method was used. The scheme, shown in Table 1 is a modification of the scheme suggested by Salomons & Forstner (1980). One must be aware, however, that sequential extraction is only a tool to better understand the distribution of an element over the different major sediment fractions as defined in Table 1. BCR of the EC is preparing reference materials in order to express sequential extraction results relative to these standards. Meanwhile, we have tested in detail the reproducibility of the various extraction steps included in our method, as well as the efficiency of each extraction step. These results, which will be published in a related paper (part of them have been presented by Baeyens et al., 1988), allow us to consider the sequential extraction technique as a valuable, additional tool for the interpretation of the geochemical behaviour of metals in sediments.
Results and Discussion The mud fiat has a heterogeneous composition showing fine sediment at station B and coarser sediment at station A. Sediment A has a thin dark brown surface layer composed of fine material while below coarse brown material is present. Sediment B consists of fine, dark brown material except for a black layer at 3-5 cm depth with a strong smell of hydrogen sulphide. 93.5% of sediment A and 39.8% of sediment B have a granulometric size < 63 ~tm. In sediment B, the redox profile (D. Van Gansbeke, personal communication) steeply drops from the first cm depth and tends to stabilize at --210 mV from the depth of 6 cm on, indicating a shallow switch from an oxic to an anoxic condition (Table 2). In contrast the oxic/anoxic interface is situated at about 8-10 cm depth at station A. The ignition loss profiles of the bulk sediments are shown in Fig. 2; they vary from 1.36 to 3.52% in sediment A and from 4.30 to 10.37% in B. The metal contents at stations A and B in the bulk sediment and in the fraction < 63 ~tm are presented in Figs 2 and 3, respectively. They are qualitatively and quantitatively different between both stations, but at a given station all metals show a similar trend. In sediment A, the metal content increases with increasing depth whereas maximum metal contents at about 3-5 cm depth are observed in sediment B. Cd and Pb showed high levels, 41 gg g-i and 200 ~tg g-i respectively, but Baeyens et al., (1982) reported 19 ~tg g-~ Cd and 174 ~tg g-~ Pb in surface sediments collected in the navigation channel next to the Ballasplaat, and Wollast et al. (1985) reported values of 15 ~tg g-1 Cd and 455 ~tg g-~ Pb in bottom sediments between 50 and 80 km from the fiver mouth of the Scheldt estuary. Obviously, the areas where high metal contents in the Scheldt are reported are situated mainly upstream from the border between Belgium and the Netherlands (Baeyens et al., 1982; Wollast et al., 1985; this work). Downstream of 130
TABLE 2 Dissolved Cd, Pb, and redox potential in the sediments at stations A and B.
Depth
(cm)
Cd (ggl -t)
Station A Pb 0tgl -l)
Eh* (mV)
Cd (~tgl-')
Station Pb (~tgl -~)
2.4 3.4 3.2
+330 +300 +268 +254 +241 +238 +72
2.07 ND ND ND ND ND ND
2.0 1.1 1.0 3.0 2.0 2.2 2.0
B
Eh* (mV)
Surface
water 0-1 1-2 2-3 3-4 4-5 5-7.5 7.5-10 10-15 15-20 20-25
11.25 ND
ND
-
-
- - 5 8
--190 --242
-
ND ND ND
-
+311 --68 --139 --181 --205 --209 --210 --215
5.9 2.8 5.1
ND = Not detectable, - = No Data. *D. Van Gansbeke, personal communication.
this zone, the contamination in the bottom sediments is much less (Araujo et al., 1988). The sequential extraction results at both stations are plotted in Fig. 4. It appears from these results that for Mn the carbonate and exchangeable fraction is a very important one (about 30% in sediment A, 40% in sediment B). None of the other studied metals shows such a high labile sediment fraction. The reducible fraction is relatively high in the oxic layer of sediment A, but decreases rapidly in the deeper anoxic layers. In addition enrichment factors indicate a depletion of Mn in the suboxic sediment (oxygen is here exhausted but sulphate not yet reduced) and enrichment of Mn in the anoxic sediment (this is the sulphate reduction zone). Fe does not show the same behaviour as Mn, since it is mainly associated with the residual fraction (7080%). 3-8% and 5-12% of Fe was found in the exchangeable and carbonate fraction at station A and B, respectively; Fe in this fraction tends to increase with increasing depth. The reducible fraction of Fe is higher in the surface layer (the oxidized Fe species are found in the more oxic surface layers) and lower in the deeper sediment. The acid soluble Fe is higher at all depths in sediment B than A. At both stations low amounts of Fe are found in oxidizable fraction (2-3%). According to the low amount of Fe in the oxidizable fraction, pyrite is not an important fraction of Fe in this mud flat, not even in the anoxic sediment layers. This is in contrast with Lord & Church (1983) who found a large accumulation of pyrite at the O2/H2S interface (above this zone little pyrite accumulated due to extensive reoxidation). The acid soluble Fe is more important than the pyrite fraction in our sediments A and B. In the surface and suboxic sediment layers of sediments A and B, a substantial fraction of the Cd content was found to be associated with the exchangeable and carbonate fraction and the reducible fraction: in sediment A 40-45%, in sediment B about 15%. However, most of the Cd content at both stations was found in the oxidizable fraction, which is theoretically composed of organic-metal and metal-sulphide complexes, showing an increase with increasing depth and decreasing redox potential. Jacobs & Emerson (1982) concluded in their
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work that heavy metals such as Cu, Zn, and Cd display a dramatic solubility decrease across the oxic/anoxic
interface leading to enrichment in the anoxic sediment layers. In our study, no dissolved Cd was observed in the pore waters. This might be a consequence of a very fast formation of a strong cadmium sulphide complex (Emerson et al., 1983; Boulegue, 1983). The enrichment factor for Cd is about 10 times higher in sediment B than A, and is most pronounced at a depth of 5-7.5 131
Marine Pollution Bulletin STATION B
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132
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Volume 22/Number 3/March 1991
cm in sediment B, where the black colour and strong smell of H2S is observed. The work of Jacobs & Emerson (1982) and Emerson et al. (1983), who studied the partitioning of metals across the O2/H2S interface in a permanently anoxic basin, supports our conclusion that metal sulphide precipitation influences the magnitude of the metal enrichment in sediments. Nevertheless, dissolved Cd was found about five times higher in the overlying water at station A, compared to station B, i.e. 11 ~tg 1-1 and 2 ~tg 1-1 respectively. This is in contrast to the Cd content in the sediments at each station as shown above, but two possible explanations exist. Station A is situated close to the discharge pipe of the BASF factory (producing fertilizers, plastics etc.) and to the discharge point of a small river, the Schijn, which is pumped into the Scheldt. The Schijn flows through the industrial area of Antwerp situated on the right bank of the river. These two outflows may have a strong influence on the local high dissolved Cd values. In addition, sampling was carried out at shallow water depth. Oxidation of reduced cadmium during low tide, may also influence locally the water column concentrations. In contrast to Cd, Pb is mainly associated with the residual fraction (about 60% at both stations). The other 40% is shared by the other fractions, with the acid soluble fraction the smallest. The exchangeable and carbonate fraction and the reducible fraction are more important in the surface layers while in the deeper sediment layers these two fractions diminish in favour of the oxidizable fraction. The size of the residual fraction is hardly affected by the redox conditions. The pore water results at station B show an increase of dissolved Pb with increasing depth. This would indicate that there is a competition between lead sulphide formation and complexation by dissolved organic ligands (e.g. humic acids and others) which control the dissolved/particulate distribution of lead. The metal contents in the fraction < 63 ~tm at station A are higher than in the bulk sediment, but at station B they both show a similar profile. Below 10 cm depth, the metal content in the fine sediment fraction is even higher at station A than at station B. Comparison of metal contents in the sediment fraction < 63 ~tm and in total suspended matter of the Scheldt (Table 3) indicates that the contamination of heavy metals in the fraction <63 p.m is comparable to the suspended matter at the same location, about 60 km from the river mouth. Metal accumulation at the mud flat was created by settling of this contaminated suspended matter. Therefore if the physicochemical conditions at the two
stations were the same, the same range of metal contents in the sediment size fraction < 63 ~m at both stations should be found. This was, however, not the case due to different physicochemical conditions.
Conclusion The contamination at the sampling sites is created by the settling of contaminated suspended matter and by local discharges. A dike, which separates the mud flat in two parts, creates different hydrodynamic conditions and hence different sediment structures and different physicochemical conditions in the sediments at each side of the dike. Station A is located in the sand and coarser grain size area, station B in the mud and finer grain size area. The upper zone of station A is oxic to suboxic while below 1 cm at station B and below 10 cm at station A the sediment is strong anoxic. In the suboxic sediment layers, Fe and Mn were reduced and redissolved but simultaneously other metals such as Pb were also resolubilized. Therefore we found less contamination in the first 10 cm at station A. At station B, which is a favourable place for deposition of fine suspended matter, a strong anoxic condition was emphasized by the imbalance between supply and demand of oxygen. It acts as a trap for metals in the Scheldt estuary due to the formation of poorly soluble metal complexes. The higher metals contamination at this station is very well supported by the observation of lower species diversity and biomass. Finally, we found that the geochemical behaviour of each of the studied metals appears to be different, leading to different sediment profiles in the pore water, the bulk sediment and the sediment fractions. The authors thank the Belgium Ministry for Development and Cooperation (ABOS) and the VUB (FAME) for a research grant for S.P. They are also grateful to F. Dehairs for his comments on the manuscript.
Alexander, M. (1977). Introduction to Soil Microbiology. John Wiley & Sons, Inc., New York. Araujo, M. F. D., Bernard, P. C. & Van Grieken, R. E. (1988). Heavy metal contamination in sediments from the Belgium Coast and Scheldt estuary. Mar. Pollut. Bull. 19, 269-273. Baeyens, W., Wartel, S., Dehairs, E, Decadt, G., Bogaert, M., Gillain, G. & Dedeurwaerder, H. (1982). The river Scheldt as a transport route for heavy metals in the sea. In Distribution, Transport and Fate of Heavy Metals in the Belgian Coastal Marine Environment (A. Disteche & I. Elskens, eds), pp. 87-108. Science Policy, Brussels. Baeyens, W., Monteny, E, Leermakers, M., Lansens, P. & Vandenhoudt, A. (1988). Study of the degree of contamination of dredged sludge in the Scheldt estuary (in Dutch). Synthesis report. University of Brussels (VUB), Brussels.
TABLE 3
Comparison of the metals concentrations in the sediment fraction < 63 p~m in the anoxic sediment layers at stations A and B, and in suspended matter at 60 km from the river mouth. Distance from the river mouth (km) 60 60 60 Station A Station B
AI (mg g-i)
Fe (mg g-i)
Mn (mg g-i)
Cd (mg g-l)
Pb (mg g-i)
57.3 44.8 45.3
55.4 46.8 39.6 38.9
1475 908 772 937
22.4 29.0 22.0 7.45 15.4
174 248 237 186.5 177.7
Reference Dehairs etal. (1986) Baeyens et al. (1982) Baeyens et al. (1988) This work This work
133
Marine Pollution Bulletin Boulegue, J. (1983). Trace Metals (Fe, Cu, Zn, Cd) in Anoxic Environment. Trace Metals in Sea Water (C. S. Wong et al., eds), pp. 563-577. NATO conference series, Plenum Press, New York and London. Byers, S. C., Mills, E. L. & Stewert, P. L. (1978). A comparison of determining organic carbon in marine sediments, with suggestions for a standard method. Hydrologia 58, 43-47. Dean, W. E., Jr. (1974). Determination of carbonate and organic matter in calcareous sediments and sedimentary rocks by loss on ignition: comparison with other methods. J. Sedim. Petrol. 44,242-248. Dehairs, E & Baeyens, W. (1986). Biogeochemical regulation of stable pollutant transfer in open sea and coastal environments. Final report of contract (ENV-766-B), EEC, Brussels. Emerson, S., Jacobs, L. & Lebo, B. (1983). The behavior of heavy metals in marine anoxic water: Solubilities at the oxygen-hydrogen sulfide interface. In Trace Metals in Sea Water (C. S. Wong et al., eds), pp. 579-608. NATO conference series, Plenum Press, New York and London. Froelich, P. N., et al. (1979). Early oxidation of organic matter in pelagic sediments of the eastern equatorial Atlantic: suboxic diagenesis. Geochim. Cosmochim. Acta 43, 1075-1090. Jacobs, L. & Emerson, S. (1982). Trace metals solubility in an anoxic fjord. Earth Planet. Sci. Lett. 60,237-252. Jorgensen, B. B. (1989). Biogeochemistry of chemoautotrophic
Marine I'olh¢tion Bulletin, Volume 22, No. 3. pp. 134-140, 1991. Printed in Great Britain.
bacteria. In Autotrophic Bacteria (H. G. Schlegel & B. Bowien, eds), pp. 117-146. Science Tech Publishers & Springer-Verlag, Madison. Lord, II1, C. J. & Church, T. M. (1983). The geochemistry of salt marshes: Sedimentary ion diffusion, sulfate reduction and pyritization. Geochim. Cosmochim. Acta 47, 1381-1391. Salomons, W. & Forsmer, U. (1980). Trace metal analysis on polluted sediments. Part II: Evaluation of environmental impact. Environ. Techno. Lett. 1,506-517. Somville, M. & De Pauw, N. (1982). Influence of temperature and river discharge on water quality of the Western Schheldt estuary. Water Res. 16, 1349-1356. Strickland, J. D. H. & Parsons, T. R. (1968). A Practical Handbook of Seawater Analysis. Fisheries Research board of Canada. Bulletin No. 167. Sundby, B., Bouchard, G., Lebel, J. & Silverberg, N. (1981). Rates of organic matter oxidation and carbon transport in early diagenesis of marine sediment. Adv. Org. Geochem. 350-354. Van Impe, J. (1985). Estuarine pollution as probable causes of increase of Estuarine birds. Mar. Pollut. Bull. 16, 271-276. Wollast, R., Devos, G. & Hoenig, M. (1985). Distribution of heavy metals in the sediment of the Scheldt estuary. In Progress in Belgian Oceanographic Research (R. van Grieken & R. Wollast, eds), pp. 147-159.
0025-326X/91 S3.00+0.00 © 1991 Pergamon Press plc
Effect of Three Primary Treatment Sewage Outfalls on Metal Concentrations in the Fish Cheilodactylusfuscus Collected Along the Coast of Sydney, Australia C. McLEAN, A. G. MISKIEWICZ and E. A. ROBERTS Scientific Services and the Environmental Projects Group, Water Board, PO Box A53, Sydney, 2000, N.S. W. Australia
Samples of muscle tissue from red morwong Cheilodactylus fuscus (Pisces: Cheilodactylidae) collected at 24 sites along the coast near Sydney Australia were analysed for total concentrations of mercury, arsenic, selenium, zinc, cadmium, lead, nickel and copper. The sites were equally spaced around each of the three major ocean disposal sewage treatment plants (STP) in Sydney located at North Head, Bondi and Malabar. The mean concentrations of mercury, arsenic and zinc were highest in fish caught off Sydney Harbour and Malabar STP. The proportion of fish in which cadmium was detected was also highest off Sydney Harbour and south of Malabar STP. Average selenium concentrations decreased from north to south with peak concentrations at each STP. Many individual values of copper, lead, and nickel were near or below the detection limits and no notable trends were observed for these metals. Mercury was the only metal whose overall mean concentration exceeded
134
the National Health and Medical Research Council (NHMRC) Maximum Residue Limit (MRL).
Worldwide concern has been expressed about the effect of sewage and industrial effluent on the marine environment and numerous studies have been undertaken to assess this. Higher levels of heavy metals such as mercury, lead, copper, zinc and cadmium, have been recorded in the water, sediments and invertebrates near discharge points than at control sites (Greig & Wenzloff, 1977; Roth & Hornung, 1977; Amiel & Navrot, 1978; Young et al., 1981; Smith et al., 1981; Talbot & Chegwidden, 1982; Vasilikiotis et al., 1983; Ward et al., 1986). However usually only mercury was found at higher levels in fish (Dix & Martin, 1975; Young, 1982; Clark & Topping, 1989).