- Email: [email protected]
Source and distribution of trace metals in the Medway and Swale estuaries, Kent, UK
226
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Australian Journal of Marine and Freshwater Research 33, 1057– 1070. US EPA, 1996. SW-846 Test methods for evaluating solid waste physical/ chemical methods. US Environmental Protection Agency, Office of Solid Waste and the National Technical Information Service, Springfield, Virginia, CD-ROM.
US EPA, 2000. National guidance. Guidance for assessing chemical contaminant data for use in fish advisories. Risk assessment and fish consumption limits, third ed., Section 4, risk-based consumption limit tables, vol. 2. Available from:.
0025-326X/$ - see front matter 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2005.10.012
Source and distribution of trace metals in the Medway and Swale estuaries, Kent, UK K.L. Spencer
a,*
, C.L. MacLeod b, A. Tuckett c, S.M. Johnson
c
a
b
Geography Department, Queen Mary, University of London, Mile End Road, London, E1 4NS, UK Arcadis Geharty and Miller International, Unit 2 Craven Court, Willie Snaith Road, Newmarket, Suffolk, CB8 7FA, UK c Department of Earth and Environmental Sciences, University of Greenwich, Chatham Maritime, Kent, ME4 4TB, UK
Millions of tonnes of fine-grained sediment are dredged annually from EuropeÕs estuaries to maintain navigation routes and permit access to commercial ports and berths. The majority of this material is removed using traditional dredging techniques and disposed of in landfill or at sea. However, the implementation of the EU Landfill Directive has increased the costs associated with the disposal of dredged material in landfill and there is a growing consensus within Europe to manage fine sediment sustainably and examine the beneficial re-use of dredged material. As a result alternative techniques for both sediment dredging and disposal are being investigated throughout Europe. A number of fine sediment management techniques are currently in operation in the Medway and Swale estuaries. An extensive traditional dredging programme maintains navigation channels and water injection dredging (WID), a relatively novel dredging technique in the UK, is being used to enable continued access to industrial facilities. In addition, English Nature is keen to support a strategy of beneficial sediment re-use within the Medway, using dredged sediment to re-charge or re-create inter-tidal habitats that are currently under threat from sea level rise and associated sediment loss (English Nature, 1999). Water injection dredging involves the injection of jets of water into bottom sediments in situ, decreasing the cohesion between sediment particles and creating a turbulent water–sediment mixture. This mixture acts as a fluid with extremely low viscosity and is able to flow under the influ-
*
Corresponding author. Tel.: +44 (0) 207 882 7814; fax: +44 (0) 208 981 6276. E-mail address: [email protected] (K.L. Spencer).
ence of gravity along the sediment water interface. Environmental concerns still surrounding the use of WID are the potential release of sediment-bound contaminants to the over-lying water column and the re-distribution of contaminated sediments over wider areas (Meyer-Nehls et al., 2000; Sullivan, 2000; Ospar Commission, 2004; Spencer et al., 2005). Preliminary investigations have indicated that the Medway Estuary currently receives low but appreciable inputs of metal contaminants with numerous point and diffuse sources (Spencer, 2002) and a range of historical contaminant inputs are implied in the vertical distributions of heavy metals in salt marsh sediments (Spencer et al., 2003). However, data are yet to be presented providing a detailed picture of surface sediment quality across the estuary and no data are available for the Swale. Where sediment management practices involve the re-distribution of sediment via WID or the beneficial re-use of dredged sediment in habitat re-creation schemes it is vital that those responsible for the management of sediment have access to detailed and accurate information regarding sediment quality. In addition, the EU Water Framework Directive is likely to require the development of sediment environmental quality standards (EQS) enforceable on a pass/fail basis (Crane, 2003). Hence, in this study we have investigated the source, magnitude and distribution of metals within the surface sediments of the inter-tidal zone in the Medway and Swale estuaries and made assessments regarding sediment quality by comparison with published sediment quality guidelines. Forty five surface sediment samples were collected from the inter-tidal zone during June and July 2000 (Fig. 1). Redox potential and pH were recorded in situ, once the sediment was returned to the laboratory it was stored at 40 C until required.
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
227
Fig. 1. Sample site locations.
Moisture content is expressed as a percentage of the oven dried sediment weight and was measured by weighing 1 g (±0.001) wet sediment into a pre-dried and weighed crucible. The mass was recorded and the samples were placed in an oven at 110 C for 24 h, cooled in a dessicator and then re-weighed. The sample was then ignited at 450 C for 12 h, cooled in a dessicator and re-weighed. Loss on ignition is expressed relative to the mass of oven dried sediment weight. For sediment metal analysis 1 g (±0.001) wet sediment was digested in Teflon beakers with 10 ml Aqua Regia at 160 C for 24 h. The digestate was filtered through 0.45 lm glass fibre filter paper and the residue washed with de-ionised water. The samples were then allowed to boil to near dryness and were brought back into solution using 10 ml of 10 M HNO3. The samples were then taken to near dryness again and brought back into solution using 3% HNO3 w/v and diluted to 100 ml in a volumetric flask. The samples were analysed for a suite of metals by inductively coupled plasma emission spectrophotometry (VGL Elemental Horizon Sequential). Sediment samples for total Hg analysis were digested according to the US EPA (Method 7471B) (Environmental Protection Agency, 1998). Between 0.5 and 0.6 g wet sediment was digested in 5 ml H2SO4, 2 ml HNO3 and 5 ml KMnO4. The samples were shaken thoroughly and digested at 90 C for 90 min in a water bath. After cooling to room temperature, the sam-
ples were diluted to 100 ml with distilled water and 6 ml of 10% NH2OH. HCl was added. The sample digestates were reduced with SnCl2 in HCl and total Hg was analysed by cold vapour atomic absorption spectrometry (CETAC). For all analyses precision was monitored by analysing triplicate samples and accuracy was assessed by comparing the data to reference sediment sample determinations (LGC 6139 and LGC 6137). Accuracy was generally within ±15% of certified values (10% for Hg analysis). Percentage relative standard deviations are variable but are generally below 15% for most samples. Mean concentrations and ranges for metal data for the Medway are presented in Table 1. Surface sediments in the Medway and Swale are moderately contaminated compared to other UK estuaries (Table 2). This is in agreement with preliminary data reported by Spencer (2002). Concentrations of Cu, Hg, Pb and Zn are all elevated in the Medway and Swale compared to local geochemical background concentrations for south east England calculated by OÕReilly Wiese et al. (1995) (Table 2) while average concentrations of Ni appear to be near background. This suggests that these metals may have anthropogenic inputs to the estuary. Fig. 2 shows the distribution of Pb within the estuary and elevated concentrations are identified at samples sites 27, 35 and 56 (203 mg kg1, 102 mg kg1 and 202 mg kg1,
228
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Table 1 Concentration of metals in sediments of the Medway Estuary, UK Average
Range
%(DW) Al Ca Fe Mg
11.4 0.6 2.3 3.5
1.3–27.2 0.1–0.9 0.7–7.7 1.2–5.8
mg kg DW1 Ba Co Cu Li Mn Ni Pb Ti V Zn Zr
52 8 32 17 523 23 43 178 46 106 4
9–296 3–16 9–103 2–39 219–1878 4–70 8–203 20–302 13–82 20–392 nd–11
lg kg DW1 Hg
305
19–1302
respectively) compared to an estuary average of 43 mg kg1. After geochemical normalisation these hotspots are still apparent, hence this distribution is not dependant upon grain size variation within the estuary. The distributions of Cu, Ni and Zn within the estuary also show a similar pattern. The highest concentrations of Cu (103 mg kg1 compared to an estuary average of 32 mg kg1) and Ni (70 mg kg1 compared with an estuary average of 23 mg kg1) are both found at site 56 while the highest concentration of Zn (392 mg kg1 compared with an estuary average of 106 mg kg1) is found at site 27. This may suggest that these three locations are major source areas for a suite of metals within the estuary. Sample site 27, at Stoke Marshes on the north side of the estuary, is in proximity to an industrial waste landfill site and Kingsnorth Power Station (coal powered). A number of organic and inorganic chemical industries and a sewage treatment works currently discharge into the Swale Estuary adjacent
to sample site 56 at Queensborough, where there is also a landfill site taking construction waste and dredgings (Environment Agency, 2004). Sample site 35, Otterham Quay, is on the southern shore of the estuary close to Motney Hill sewage treatment works. Secondary sewage treatment has been used at Motney Hill STW since 2000 and hence prior to this the STW may have discharged a range of heavy metals to controlled waters. Additionally, Otterham houses a busy boatyard and a number of light industrial units and during the 19th and 20th centuries would have been used for the landing of a variety of industrial cargo. Therefore Pb, Cu, Ni and Zn appear to be entering the estuary associated with a variety of industrial discharges. Concentrations of Ba in the estuary are not particularly elevated however on closer inspection Ba has a similar distribution to Cu, Pb, Zn and Ni with elevated concentrations at site 56 (Queensborough, 296 mg kg1) and to a lesser extent at site 20 (Bloors Wharf, 153 mg kg1). Reemtsma et al. (2000) have identified urban run-off as a potential source of Ba to the estuarine environment and this is a potential source of Ba in these locations. Mercury shows a very different distribution within the estuary (Fig. 3) with the main areas of elevated concentration on the southern side of the estuary between Motney Hill and Horrid Hill at sites 19, 20 and 21 (838 lg kg1, 1302 lg kg1 and 823 lg kg1 respectively compared to an estuary average of 305 lg kg1). This distribution suggests that Hg has different sources within the estuary compared to the other metals. Important inputs of Hg to aquatic environments include atmospheric deposition and fossil fuel combustion (Carpi, 1997; Wang et al., 2004). Mercury contamination is also frequently associated with paper manufacture as phenylmercuric acetate was used as a slimicide in wood storage piles up until the middle of the 20th century (Bru¨chert, 1998). Paper manufacture has been an important industry in the Medway since c. 1700 with numerous paper mills situated upstream and on the banks of the estuary itself (Penn, 1981) and may have therefore supplied a historic source of Hg to the estuary. These sources contribute to the elevated Hg concentrations
Table 2 Comparative mean metal concentrations (lg g1) in the Medway and other UK estuaries
Cu Hg Ni Pb Zn V
Humbera
Solwayb
Forthc
Thamesd
Medway (this study)
Estimated natural backgrounde
70 NA 55 127 319 209
7 NA 17 25 59 NA
86 0.4 NA 89 150 NA
24 0.2 21 63 115 49
32 0.3 23 43 106 46
20 <0.1 30 20 50 NA
NA—data unavailable. a Grant and Middleton (1990). b Bryan and Langston (1992). c Lindsay and Bell (1997). d Attrill and Thomes (1995). e OÕReilly Wiese et al. (1995).
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Fig. 2. Distribution of Pb in surface sediments in mg kg DW1.
Fig. 3. Distribution of Hg in surface sediments in lg kg DW1.
229
230
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Table 3 Comparison of sediment quality guidelines with average values for the Medway Estuary Sediment metal concentrations in lg g1 (except Hg in lg kg1) ERL (NOAA)a,b ERM (NOAA)a,b Average Medway sediment % Samples which exceed ERL % Samples which exceed ERM
Cu
Hg
Ni
Pb
Zn
34 270 32 41 0
0.15 0.71 0.30 69 8
21 52 23 51 5
47 218 43 33 0
150 410 106 13 10
a The ERL is the concentration below which adverse effects are known to be absent in the majority of examples (usually 90%), while the ERM is the concentration above which adverse effects occur in more than 50% of cases (Environment Agency, 2002). b Values from Buchman (1999).
above local geochemical background across the estuary. Bloors Wharf, a disused boat yard, lies between Horrid Hill and Motney Hill and was previously used for boat scrapping and this may therefore present a potential point source discharge of Hg to the sediments which accounts for the contaminant hotspots in this area. Sediment quality in the study area was assessed by comparing the data to published sediment quality guidelines. Currently, sediment quality guidelines do not exist under UK legislation, however, the Environment Agency of England and Wales favour an effects range approach such as that used in the USA (Environment Agency, 2002). Here, Table 3 compares sediment metal concentrations in the Medway and Swale with effects range low (ERL) and effects range median (ERM) guidelines for marine sediments for selected metals using SQuiRTs (sediment screening reference tables published by the National Oceanographic and Atmospheric Administration). Over 69% of sample sites within the Medway and Swale will exceed the ERL for Hg and 51% for Ni. Sediment metal concentrations in excess of the ERL indicate that there may be some ecotoxicological risk to organisms living in the sediment and under the US system this warrants further investigation. This suggests that the implementation of EQSs under the Water Framework Directive may be problematic for the Medway and Swale with large areas of the study site in breach of EQSs and hence unsuitable for WID or for use in beneficial sediment reuse schemes. Acknowledgements S.M. Johnson would like to thank EPSRC for a research studentship and A. Tuckett would like to thank the University of Greenwich for a student bursary and the Medway Ports Authority for assistance with sample collection. References Attrill, M.J., Thomes, R.M., 1995. Heavy metal concentrations in sediment from the Thames estuary, UK. Marine Pollution Bulletin 30 (11), 742–744.
Bru¨chert, V., 1998. Early diagenesis of sulfur in estuarine sediments: the role of sedimentary humic and fulvic acids. Geochimica et Cosmochimica Acta 62 (9), 1567–1586. Bryan, G.W., Langston, W.J., 1992. Bioavailability, accumulation and effects of heavy metals in sediments with special reference to United Kingdom estuaries: a review. Environmental Pollution 76, 89– 131. Buchman, M.F., 1999. NOAA screening quick reference tables. NOAA HAZMAT Report 99-1, Seattle, WA, Coastal Protection and Restoration Division, National Oceanic and Atmospheric Administration, p. 12. Carpi, A., 1997. Mercury from combustion sources: a review of the chemical species emitted and their transport in the atmosphere. Water, Air and Soil Pollution 98, 241–254. Crane, M., 2003. Proposed development of sediment quality guidelines under the European Water Framework Directive: a critique. Toxicology Letters 142, 195–206. English Nature, 1999. United Kingdom Biodiversity Action Group. Tranche 2. Actions plans, vol. 5. Environment Agency, 2002. Review and recommendations of methodologies for the derivation of sediment quality guidelines. R&D Technical Report P2-082/TR. Environment Agency, 2004. WhatÕs in your backyard. Available from:. Grant, A., Middleton, R., 1990. An assessment of metal contamination in the sediments of the Humber Estuary, UK. Estuarine Coastal and Shelf Science 31, 71–85. Lindsay, P., Bell, F.G., 1997. Contaminated sediment in two United Kingdom estuaries. Environmental and Engineering Geoscience 3 (3), 375–387. Meyer-Nehls, R., von Gabriele Go¨nnert, B., Christiansen, H., Rahlf, H., 2000. Das Wasserinjektionsverfahren: Ergebnisse einer literaturstudie sowie von untersuchungen im Hamburger Hafen und in der Unterelbe. Ergebnisse aus dem Baggergutuntersuchungsprogramm. Heft 8. OÕReilly Wiese, S.B., Bubb, J.M., Lester, J.N., 1995. The significance of sediment metal concentrations in two eroding Essex saltmarshes. Marine Pollution Bulletin 30 (3), 190–199. Ospar Commission, 2004. Environmental impacts to marine species and habitats of dredging for navigational purposes. Biodiversity Series. ISBN 1-904426-50-6. Penn, R., 1981. Portrait of the River Medway. Robert Hale Ltd., London. Reemtsma, T., Gnirss, R., Jekel, M., 2000. Infiltration of combined sewer overflow and tertiary treated municipal wastewater: an integrated laboratory and field study on various metals. Water Environment Research 72 (6), 644–650. Spencer, K.L., 2002. Spatial distribution of metals in the inter-tidal sediments of the Medway Estuary, Kent. Marine Pollution Bulletin 44 (9), 933–944.
Baseline / Marine Pollution Bulletin 52 (2006) 214–238 Spencer, K.L., Cundy, A.B., Croudace, I.W., 2003. Heavy metal distribution and early-diagenesis in salt marsh sediments from the Medway Estuary, Kent, UK. Estuarine Coastal and Shelf Science 57 (1–2), 43–54. Spencer, K.L., Dewhurst, R.E., Penna, P., 2005. Potential impacts of water injection dredging on water quality and ecotoxicity in Limehouse Basin. Chemosphere.
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Sullivan, N., 2000. The use of agitation dredging, water injection dredging and sidecasting: result of a survey of ports in England and Wales. Terra et Aqua 78, 11–20. Wang, Q.R., Kim, D., Dionysiou, D.D., Sorial, G.A., Timberlake, D., 2004. Sources and remediation for mercury contamination in aquatic systems—a literature review. Environmental Pollution 131 (2), 323– 336.
0025-326X/$ - see front matter 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2005.10.019
A survey of organic solvent extractable metal concentrations in the bottom sediments in Osaka Bay, Japan Mika Yamada a
a,*
, Koshi Yamamoto a, Yasuhiro Ushihara b, Hiroshi Kawai
b
Department of Earth and Planetary Sciences, Graduate School of Environmental Studies, Nagoya University, Furo, Chikusa, Nagoya, 464-8602, Japan b Kobe University Research Center for Inland Seas, 1-1 Rokkodai, Nada, Kobe, 657-8501, Japan
Metal elements contained in marine sediments occur in a variety of chemical forms, which can be broadly classified into inorganic and organic. These two forms of metals exhibit very different effects on the metabolism of various organisms (Crompton, 1998). In general, organo-metal compounds such as -Hg and -Sn are more toxic than the inorgano-metals from which they are derived (Craig et al., 2003). A notable exception is organo-arsenic, which is markedly less toxic than inorgano-As for a wide variety of organisms (Edmonds and Francesconi, 2003). Metal elements are an important component of coastal pollutants (Pichtel et al., 1997; Imperato et al., 2003). However, in assessing the influence of these elements on the marine environment, it is essential to determine whether they are in organic or inorganic forms. In this study of organo- and inorgano-metal compositions in marine sediments, we extracted metals from sediment samples collected in Osaka bay using an organic solvent and determined metal concentrations in the extracts. We refer to the metals extracted in this way as ‘‘organic solvent extractable (OSE) metals’’. Osaka Bay is located on the eastern most part of the Seto Inland Sea, the largest enclosed coastal sea in Japan. The bay is bounded by Honshu and Awaji Islands and is connected to the Pacific Ocean through the Kitan Straits and to the central part of the Seto Inland Sea through the Akashi Straits (Fig. 1). The current bay morphology can be broadly divided into two parts by the 20 m isobath (Nakatuji, 1998): the area on the NE side is shallower than 20 m (SA in Fig. 1) and the area on the SW side (AW, DA * Corresponding author. Present address: Graduate School of Science and Technology, Kobe University, 1-1 Rokodai, Nada, Kobe, 657-8501, Japan. Tel./fax: +81 7880 35781. E-mail address: [email protected] (M. Yamada).
and IA in Fig. 1) deeper than 20 m. The former area is rather stagnant, whereas the latter area is relatively well mixed both horizontally and vertically by the strong tidal current through the straits. On the NE side of the bay lies Osaka plain. This is a highly industrialized area with a population of about 16 million. A large quantity of drainage water from domestic and industrial sources flows into the bay through the ShinYodo, Yamato and some other smaller rivers. The water flow into the bay is estimated to be 13.1 · 109 m3 per year (Hoshika et al., 2000). The NE side of the bay is, therefore, eutrophic and shows high primary productivity. The biochemical oxygen demand (BOD) of the area is 1.5– 370 mg/L and the chemical oxygen demand (COD) is 20.7–27.7 mg/g (Hyogo Prefecture, 2003). In contrast, the seawater on the SW side of the bay facing Awaji Island is less eutrophic, because of the more frequent water exchange and fewer drainage sources. Awaji Island is much less industrialized than the Osaka Plain area and the population of the island is ca. 160,000. The BOD of the Awaji coast area is 1.4–4.5 mg/L and COD is 1–1.7 mg/g (Hyogo Prefecture, 2003). Bottom sediments of the inner part of the bay are reported to contain relatively high concentrations of heavy metal elements (Kitano et al., 1981; Gohda and Yamazaki, 1982; Sue et al., 1983; Hoshika and Shiozawa, 1986) and this is thought to be due to industrial discharge especially during the period of the rapid industrial development from the 1950s to 1970s. The remarkable contrast in the water quality between the NE and SW sides of the Bay presents an opportunity to examine a new method for assessing metal element pollution in the sediments. Bottom sediment samples were collected at 154 sites covering most of Osaka Bay (Fig. 1) using a plastic tube core (5 cm in diameter, gravity type KK, Hashimoto-Kagaku,
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Australian Journal of Marine and Freshwater Research 33, 1057– 1070. US EPA, 1996. SW-846 Test methods for evaluating solid waste physical/ chemical methods. US Environmental Protection Agency, Office of Solid Waste and the National Technical Information Service, Springfield, Virginia, CD-ROM.
US EPA, 2000. National guidance. Guidance for assessing chemical contaminant data for use in fish advisories. Risk assessment and fish consumption limits, third ed., Section 4, risk-based consumption limit tables, vol. 2. Available from:
0025-326X/$ - see front matter 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2005.10.012
Source and distribution of trace metals in the Medway and Swale estuaries, Kent, UK K.L. Spencer
a,*
, C.L. MacLeod b, A. Tuckett c, S.M. Johnson
c
a
b
Geography Department, Queen Mary, University of London, Mile End Road, London, E1 4NS, UK Arcadis Geharty and Miller International, Unit 2 Craven Court, Willie Snaith Road, Newmarket, Suffolk, CB8 7FA, UK c Department of Earth and Environmental Sciences, University of Greenwich, Chatham Maritime, Kent, ME4 4TB, UK
Millions of tonnes of fine-grained sediment are dredged annually from EuropeÕs estuaries to maintain navigation routes and permit access to commercial ports and berths. The majority of this material is removed using traditional dredging techniques and disposed of in landfill or at sea. However, the implementation of the EU Landfill Directive has increased the costs associated with the disposal of dredged material in landfill and there is a growing consensus within Europe to manage fine sediment sustainably and examine the beneficial re-use of dredged material. As a result alternative techniques for both sediment dredging and disposal are being investigated throughout Europe. A number of fine sediment management techniques are currently in operation in the Medway and Swale estuaries. An extensive traditional dredging programme maintains navigation channels and water injection dredging (WID), a relatively novel dredging technique in the UK, is being used to enable continued access to industrial facilities. In addition, English Nature is keen to support a strategy of beneficial sediment re-use within the Medway, using dredged sediment to re-charge or re-create inter-tidal habitats that are currently under threat from sea level rise and associated sediment loss (English Nature, 1999). Water injection dredging involves the injection of jets of water into bottom sediments in situ, decreasing the cohesion between sediment particles and creating a turbulent water–sediment mixture. This mixture acts as a fluid with extremely low viscosity and is able to flow under the influ-
*
Corresponding author. Tel.: +44 (0) 207 882 7814; fax: +44 (0) 208 981 6276. E-mail address: [email protected] (K.L. Spencer).
ence of gravity along the sediment water interface. Environmental concerns still surrounding the use of WID are the potential release of sediment-bound contaminants to the over-lying water column and the re-distribution of contaminated sediments over wider areas (Meyer-Nehls et al., 2000; Sullivan, 2000; Ospar Commission, 2004; Spencer et al., 2005). Preliminary investigations have indicated that the Medway Estuary currently receives low but appreciable inputs of metal contaminants with numerous point and diffuse sources (Spencer, 2002) and a range of historical contaminant inputs are implied in the vertical distributions of heavy metals in salt marsh sediments (Spencer et al., 2003). However, data are yet to be presented providing a detailed picture of surface sediment quality across the estuary and no data are available for the Swale. Where sediment management practices involve the re-distribution of sediment via WID or the beneficial re-use of dredged sediment in habitat re-creation schemes it is vital that those responsible for the management of sediment have access to detailed and accurate information regarding sediment quality. In addition, the EU Water Framework Directive is likely to require the development of sediment environmental quality standards (EQS) enforceable on a pass/fail basis (Crane, 2003). Hence, in this study we have investigated the source, magnitude and distribution of metals within the surface sediments of the inter-tidal zone in the Medway and Swale estuaries and made assessments regarding sediment quality by comparison with published sediment quality guidelines. Forty five surface sediment samples were collected from the inter-tidal zone during June and July 2000 (Fig. 1). Redox potential and pH were recorded in situ, once the sediment was returned to the laboratory it was stored at 40 C until required.
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
227
Fig. 1. Sample site locations.
Moisture content is expressed as a percentage of the oven dried sediment weight and was measured by weighing 1 g (±0.001) wet sediment into a pre-dried and weighed crucible. The mass was recorded and the samples were placed in an oven at 110 C for 24 h, cooled in a dessicator and then re-weighed. The sample was then ignited at 450 C for 12 h, cooled in a dessicator and re-weighed. Loss on ignition is expressed relative to the mass of oven dried sediment weight. For sediment metal analysis 1 g (±0.001) wet sediment was digested in Teflon beakers with 10 ml Aqua Regia at 160 C for 24 h. The digestate was filtered through 0.45 lm glass fibre filter paper and the residue washed with de-ionised water. The samples were then allowed to boil to near dryness and were brought back into solution using 10 ml of 10 M HNO3. The samples were then taken to near dryness again and brought back into solution using 3% HNO3 w/v and diluted to 100 ml in a volumetric flask. The samples were analysed for a suite of metals by inductively coupled plasma emission spectrophotometry (VGL Elemental Horizon Sequential). Sediment samples for total Hg analysis were digested according to the US EPA (Method 7471B) (Environmental Protection Agency, 1998). Between 0.5 and 0.6 g wet sediment was digested in 5 ml H2SO4, 2 ml HNO3 and 5 ml KMnO4. The samples were shaken thoroughly and digested at 90 C for 90 min in a water bath. After cooling to room temperature, the sam-
ples were diluted to 100 ml with distilled water and 6 ml of 10% NH2OH. HCl was added. The sample digestates were reduced with SnCl2 in HCl and total Hg was analysed by cold vapour atomic absorption spectrometry (CETAC). For all analyses precision was monitored by analysing triplicate samples and accuracy was assessed by comparing the data to reference sediment sample determinations (LGC 6139 and LGC 6137). Accuracy was generally within ±15% of certified values (10% for Hg analysis). Percentage relative standard deviations are variable but are generally below 15% for most samples. Mean concentrations and ranges for metal data for the Medway are presented in Table 1. Surface sediments in the Medway and Swale are moderately contaminated compared to other UK estuaries (Table 2). This is in agreement with preliminary data reported by Spencer (2002). Concentrations of Cu, Hg, Pb and Zn are all elevated in the Medway and Swale compared to local geochemical background concentrations for south east England calculated by OÕReilly Wiese et al. (1995) (Table 2) while average concentrations of Ni appear to be near background. This suggests that these metals may have anthropogenic inputs to the estuary. Fig. 2 shows the distribution of Pb within the estuary and elevated concentrations are identified at samples sites 27, 35 and 56 (203 mg kg1, 102 mg kg1 and 202 mg kg1,
228
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Table 1 Concentration of metals in sediments of the Medway Estuary, UK Average
Range
%(DW) Al Ca Fe Mg
11.4 0.6 2.3 3.5
1.3–27.2 0.1–0.9 0.7–7.7 1.2–5.8
mg kg DW1 Ba Co Cu Li Mn Ni Pb Ti V Zn Zr
52 8 32 17 523 23 43 178 46 106 4
9–296 3–16 9–103 2–39 219–1878 4–70 8–203 20–302 13–82 20–392 nd–11
lg kg DW1 Hg
305
19–1302
respectively) compared to an estuary average of 43 mg kg1. After geochemical normalisation these hotspots are still apparent, hence this distribution is not dependant upon grain size variation within the estuary. The distributions of Cu, Ni and Zn within the estuary also show a similar pattern. The highest concentrations of Cu (103 mg kg1 compared to an estuary average of 32 mg kg1) and Ni (70 mg kg1 compared with an estuary average of 23 mg kg1) are both found at site 56 while the highest concentration of Zn (392 mg kg1 compared with an estuary average of 106 mg kg1) is found at site 27. This may suggest that these three locations are major source areas for a suite of metals within the estuary. Sample site 27, at Stoke Marshes on the north side of the estuary, is in proximity to an industrial waste landfill site and Kingsnorth Power Station (coal powered). A number of organic and inorganic chemical industries and a sewage treatment works currently discharge into the Swale Estuary adjacent
to sample site 56 at Queensborough, where there is also a landfill site taking construction waste and dredgings (Environment Agency, 2004). Sample site 35, Otterham Quay, is on the southern shore of the estuary close to Motney Hill sewage treatment works. Secondary sewage treatment has been used at Motney Hill STW since 2000 and hence prior to this the STW may have discharged a range of heavy metals to controlled waters. Additionally, Otterham houses a busy boatyard and a number of light industrial units and during the 19th and 20th centuries would have been used for the landing of a variety of industrial cargo. Therefore Pb, Cu, Ni and Zn appear to be entering the estuary associated with a variety of industrial discharges. Concentrations of Ba in the estuary are not particularly elevated however on closer inspection Ba has a similar distribution to Cu, Pb, Zn and Ni with elevated concentrations at site 56 (Queensborough, 296 mg kg1) and to a lesser extent at site 20 (Bloors Wharf, 153 mg kg1). Reemtsma et al. (2000) have identified urban run-off as a potential source of Ba to the estuarine environment and this is a potential source of Ba in these locations. Mercury shows a very different distribution within the estuary (Fig. 3) with the main areas of elevated concentration on the southern side of the estuary between Motney Hill and Horrid Hill at sites 19, 20 and 21 (838 lg kg1, 1302 lg kg1 and 823 lg kg1 respectively compared to an estuary average of 305 lg kg1). This distribution suggests that Hg has different sources within the estuary compared to the other metals. Important inputs of Hg to aquatic environments include atmospheric deposition and fossil fuel combustion (Carpi, 1997; Wang et al., 2004). Mercury contamination is also frequently associated with paper manufacture as phenylmercuric acetate was used as a slimicide in wood storage piles up until the middle of the 20th century (Bru¨chert, 1998). Paper manufacture has been an important industry in the Medway since c. 1700 with numerous paper mills situated upstream and on the banks of the estuary itself (Penn, 1981) and may have therefore supplied a historic source of Hg to the estuary. These sources contribute to the elevated Hg concentrations
Table 2 Comparative mean metal concentrations (lg g1) in the Medway and other UK estuaries
Cu Hg Ni Pb Zn V
Humbera
Solwayb
Forthc
Thamesd
Medway (this study)
Estimated natural backgrounde
70 NA 55 127 319 209
7 NA 17 25 59 NA
86 0.4 NA 89 150 NA
24 0.2 21 63 115 49
32 0.3 23 43 106 46
20 <0.1 30 20 50 NA
NA—data unavailable. a Grant and Middleton (1990). b Bryan and Langston (1992). c Lindsay and Bell (1997). d Attrill and Thomes (1995). e OÕReilly Wiese et al. (1995).
Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Fig. 2. Distribution of Pb in surface sediments in mg kg DW1.
Fig. 3. Distribution of Hg in surface sediments in lg kg DW1.
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Baseline / Marine Pollution Bulletin 52 (2006) 214–238
Table 3 Comparison of sediment quality guidelines with average values for the Medway Estuary Sediment metal concentrations in lg g1 (except Hg in lg kg1) ERL (NOAA)a,b ERM (NOAA)a,b Average Medway sediment % Samples which exceed ERL % Samples which exceed ERM
Cu
Hg
Ni
Pb
Zn
34 270 32 41 0
0.15 0.71 0.30 69 8
21 52 23 51 5
47 218 43 33 0
150 410 106 13 10
a The ERL is the concentration below which adverse effects are known to be absent in the majority of examples (usually 90%), while the ERM is the concentration above which adverse effects occur in more than 50% of cases (Environment Agency, 2002). b Values from Buchman (1999).
above local geochemical background across the estuary. Bloors Wharf, a disused boat yard, lies between Horrid Hill and Motney Hill and was previously used for boat scrapping and this may therefore present a potential point source discharge of Hg to the sediments which accounts for the contaminant hotspots in this area. Sediment quality in the study area was assessed by comparing the data to published sediment quality guidelines. Currently, sediment quality guidelines do not exist under UK legislation, however, the Environment Agency of England and Wales favour an effects range approach such as that used in the USA (Environment Agency, 2002). Here, Table 3 compares sediment metal concentrations in the Medway and Swale with effects range low (ERL) and effects range median (ERM) guidelines for marine sediments for selected metals using SQuiRTs (sediment screening reference tables published by the National Oceanographic and Atmospheric Administration). Over 69% of sample sites within the Medway and Swale will exceed the ERL for Hg and 51% for Ni. Sediment metal concentrations in excess of the ERL indicate that there may be some ecotoxicological risk to organisms living in the sediment and under the US system this warrants further investigation. This suggests that the implementation of EQSs under the Water Framework Directive may be problematic for the Medway and Swale with large areas of the study site in breach of EQSs and hence unsuitable for WID or for use in beneficial sediment reuse schemes. Acknowledgements S.M. Johnson would like to thank EPSRC for a research studentship and A. Tuckett would like to thank the University of Greenwich for a student bursary and the Medway Ports Authority for assistance with sample collection. References Attrill, M.J., Thomes, R.M., 1995. Heavy metal concentrations in sediment from the Thames estuary, UK. Marine Pollution Bulletin 30 (11), 742–744.
Bru¨chert, V., 1998. Early diagenesis of sulfur in estuarine sediments: the role of sedimentary humic and fulvic acids. Geochimica et Cosmochimica Acta 62 (9), 1567–1586. Bryan, G.W., Langston, W.J., 1992. Bioavailability, accumulation and effects of heavy metals in sediments with special reference to United Kingdom estuaries: a review. Environmental Pollution 76, 89– 131. Buchman, M.F., 1999. NOAA screening quick reference tables. NOAA HAZMAT Report 99-1, Seattle, WA, Coastal Protection and Restoration Division, National Oceanic and Atmospheric Administration, p. 12. Carpi, A., 1997. Mercury from combustion sources: a review of the chemical species emitted and their transport in the atmosphere. Water, Air and Soil Pollution 98, 241–254. Crane, M., 2003. Proposed development of sediment quality guidelines under the European Water Framework Directive: a critique. Toxicology Letters 142, 195–206. English Nature, 1999. United Kingdom Biodiversity Action Group. Tranche 2. Actions plans, vol. 5. Environment Agency, 2002. Review and recommendations of methodologies for the derivation of sediment quality guidelines. R&D Technical Report P2-082/TR. Environment Agency, 2004. WhatÕs in your backyard. Available from:
Baseline / Marine Pollution Bulletin 52 (2006) 214–238 Spencer, K.L., Cundy, A.B., Croudace, I.W., 2003. Heavy metal distribution and early-diagenesis in salt marsh sediments from the Medway Estuary, Kent, UK. Estuarine Coastal and Shelf Science 57 (1–2), 43–54. Spencer, K.L., Dewhurst, R.E., Penna, P., 2005. Potential impacts of water injection dredging on water quality and ecotoxicity in Limehouse Basin. Chemosphere.
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Sullivan, N., 2000. The use of agitation dredging, water injection dredging and sidecasting: result of a survey of ports in England and Wales. Terra et Aqua 78, 11–20. Wang, Q.R., Kim, D., Dionysiou, D.D., Sorial, G.A., Timberlake, D., 2004. Sources and remediation for mercury contamination in aquatic systems—a literature review. Environmental Pollution 131 (2), 323– 336.
0025-326X/$ - see front matter 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2005.10.019
A survey of organic solvent extractable metal concentrations in the bottom sediments in Osaka Bay, Japan Mika Yamada a
a,*
, Koshi Yamamoto a, Yasuhiro Ushihara b, Hiroshi Kawai
b
Department of Earth and Planetary Sciences, Graduate School of Environmental Studies, Nagoya University, Furo, Chikusa, Nagoya, 464-8602, Japan b Kobe University Research Center for Inland Seas, 1-1 Rokkodai, Nada, Kobe, 657-8501, Japan
Metal elements contained in marine sediments occur in a variety of chemical forms, which can be broadly classified into inorganic and organic. These two forms of metals exhibit very different effects on the metabolism of various organisms (Crompton, 1998). In general, organo-metal compounds such as -Hg and -Sn are more toxic than the inorgano-metals from which they are derived (Craig et al., 2003). A notable exception is organo-arsenic, which is markedly less toxic than inorgano-As for a wide variety of organisms (Edmonds and Francesconi, 2003). Metal elements are an important component of coastal pollutants (Pichtel et al., 1997; Imperato et al., 2003). However, in assessing the influence of these elements on the marine environment, it is essential to determine whether they are in organic or inorganic forms. In this study of organo- and inorgano-metal compositions in marine sediments, we extracted metals from sediment samples collected in Osaka bay using an organic solvent and determined metal concentrations in the extracts. We refer to the metals extracted in this way as ‘‘organic solvent extractable (OSE) metals’’. Osaka Bay is located on the eastern most part of the Seto Inland Sea, the largest enclosed coastal sea in Japan. The bay is bounded by Honshu and Awaji Islands and is connected to the Pacific Ocean through the Kitan Straits and to the central part of the Seto Inland Sea through the Akashi Straits (Fig. 1). The current bay morphology can be broadly divided into two parts by the 20 m isobath (Nakatuji, 1998): the area on the NE side is shallower than 20 m (SA in Fig. 1) and the area on the SW side (AW, DA * Corresponding author. Present address: Graduate School of Science and Technology, Kobe University, 1-1 Rokodai, Nada, Kobe, 657-8501, Japan. Tel./fax: +81 7880 35781. E-mail address: [email protected] (M. Yamada).
and IA in Fig. 1) deeper than 20 m. The former area is rather stagnant, whereas the latter area is relatively well mixed both horizontally and vertically by the strong tidal current through the straits. On the NE side of the bay lies Osaka plain. This is a highly industrialized area with a population of about 16 million. A large quantity of drainage water from domestic and industrial sources flows into the bay through the ShinYodo, Yamato and some other smaller rivers. The water flow into the bay is estimated to be 13.1 · 109 m3 per year (Hoshika et al., 2000). The NE side of the bay is, therefore, eutrophic and shows high primary productivity. The biochemical oxygen demand (BOD) of the area is 1.5– 370 mg/L and the chemical oxygen demand (COD) is 20.7–27.7 mg/g (Hyogo Prefecture, 2003). In contrast, the seawater on the SW side of the bay facing Awaji Island is less eutrophic, because of the more frequent water exchange and fewer drainage sources. Awaji Island is much less industrialized than the Osaka Plain area and the population of the island is ca. 160,000. The BOD of the Awaji coast area is 1.4–4.5 mg/L and COD is 1–1.7 mg/g (Hyogo Prefecture, 2003). Bottom sediments of the inner part of the bay are reported to contain relatively high concentrations of heavy metal elements (Kitano et al., 1981; Gohda and Yamazaki, 1982; Sue et al., 1983; Hoshika and Shiozawa, 1986) and this is thought to be due to industrial discharge especially during the period of the rapid industrial development from the 1950s to 1970s. The remarkable contrast in the water quality between the NE and SW sides of the Bay presents an opportunity to examine a new method for assessing metal element pollution in the sediments. Bottom sediment samples were collected at 154 sites covering most of Osaka Bay (Fig. 1) using a plastic tube core (5 cm in diameter, gravity type KK, Hashimoto-Kagaku,