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Edited by Leon P. Zann
Great Barrier Reef Marine Park Authority, Townsville Queensland
Ocean Rescue 2000 Program
Department of the Environment, Sport and Territories, Canberra, 1995
ISBN 0 642 17406 7
CSIRO Centre for Advanced Analytical Chemistry
Menai, NSW 2234
Heavy metals are a major anthropogenic contaminant of estuarine and coastal waters. Their inputs include urban run-off, industrial effluents, mining operations and atmospheric depositions, and may be in particulate or dissolved forms. Although many are essential biological elements, all have the potential to be toxic to organisms above certain threshold concentrations, and for the protection of aquatic biota it is important that these limits not be exceeded in aquatic environments.
The oceans are often mistakenly seen as a boundless sink for heavy metals and other contaminants. In shallow near-shore waters where dispersion and dilution processes are less effective, the measured concentrations of heavy metals are generally noticeably higher than in open ocean waters. Australian coastal waters are generally characterised by currents which follow the coastline and are less readily able to transport contaminants to deeper waters. In deeper waters, metal concentrations are extremely low, and there are noticeable differences between metal concentrations in surface and bottom waters, the profiles often correlating with nutrient concentrations and organic particulate matter. In near-shore waters, however, such differences are less pronounced.
Although some metals such as molybdenum and uranium are highly conserved in marine waters, most are not, attaching to suspended particulates, and ultimately accumulating in bottom sediments. The sediment load of coastal waters is noticeably higher than open ocean waters. This is particularly so in the estuarine zone where, with increasing salinity the precipitation of iron hydrous oxides will scavenge and co-precipitate soluble metal species.
The bio-availability of both dissolved and particulate heavy metals is critically dependent on chemical form, and a great deal of research, much of it pioneered in Australia, has been devoted to methods which can detect specific species classes whose biological impacts can be predicted (Florence & Batley 1980). There are limited measurements of the speciation of metals in waters, but for the most part data are available only for total metal concentrations. These usually refer to the dissolved metal concentrations defined by filtration through a 0.45 µm membrane filter.
A number of texts have discussed the sources of heavy metals to estuaries and coastal waters (Furness & Rainbow 1990; Salomons & Forstner 1984; Forstner & Wittman 1979). In the northern hemisphere, the higher population densities and greater urbanisation and industrialisation have led to quite high contamination of coastal waters by metals. In Australia, potential problems are restricted to less than a dozen major cities located on the coastal fringe. These include the state capitals, and a number of other highly industrialised cities, such as Wollongong and Newcastle.
Because most heavy metals tend to accumulate in sediments , their presence in the water column is usually the result of recent inputs. Metal concentrations can vary significantly over short distances and as a function of tide. Single measurements at a given site may indicate contamination, but to fully assess its magnitude a more complex sampling program which considers spatial and temporal variability is required. For this, integrating samplers, which continuously preconcentrate metals from solution over periods from hours to days, have been used.
Major estuarine contamination is generally directly traceable to an industrial (or mining) source, although urban run-off can be responsible for an increase in metal concentrations above background open ocean values. In Australia, industrial and mining sources could include smelters, power stations, paper mills, port facilities including ore or coal loaders, sewage treatment works, oil refineries and other chemical and industrial manufacturing plants.
In the late 1970s, following international disasters involving mercury and cadmium, heavy metals were identified as the major pollution threat facing mankind. At that time the Derwent River in Tasmania was severely affected by discharges of metallurgical waste, partially treated sewage, and effluent from a pulp mill operation and a chlor-alkali plant (Bloom & Ayling 1977), and was described as 'one of the most polluted areas of the world' (Forstner & Wittman 1979). The past twenty years have seen a worldwide response to this threat. In Australia too, stricter controls on aqueous discharges have been progressively imposed by state and federal environmental agencies (ANZECC 1992). These have succeeded in dramatically reducing the dissolved concentrations of heavy metals in estuarine waters. Evidence for this change is available from routine monitoring data held by these agencies. More often these relate to concentrations at the point of discharge rather than the dispersed concentrations in the estuary, because firstly this is the point of regulation, and secondly the concentrations are demonstrably higher and therefore analyses are easier and more reliable. Sediments equally reflect the improvement in discharge controls, with the concentrations of heavy metals in uppermost layers being lower than those in deeper sediments. Examples of this have been found for Lake Illawarra and Lake Macquarie (NSW), as well as in offshore sediments.
Tributyltin (TBT) has been used in Australia as an active ingredient in marine antifouling paints since the early 1970s. Modern copolymer paint formulations have an initial high leach rate of TBT which, within days, reaches a standard and constant value of around 4 ug TBT cm-2 day-1. The half-life of TBT in seawater is around six hours, but it rapidly partitions either to suspended sediments or to the surface microlayer. In sediments its half-life has been estimated at around 3.5 years (Batley et al. 1992). Dissolved TBT has adverse impacts on oysters (and to a lesser extent other bivalves), with tissue growth diminished at the expense of shell, leading in some species to shell thickening or curling. These impacts led to what is effectively a worldwide ban on the use of TBT in paints on ships below 25 m in length.
In Australia, the ban took effect in Australia around 1988. TBT is however still used in paints for larger ocean-going vessels, and can therefore be detected in the waters of most major harbours, and especially in sediments in the vicinity of dockyards.
Butyltins from sources other than marine paints can also enter the water column. Dibutyltin is used as a catalyst in the plastics industry. TBT is used as an algaecide in boiler water cooling circuits. Dibutyltin is the primary degradation product of TBT, but has a comparatively minor impact on the environment. Other alkyltins have been used in pesticide formulations, but none have been detected in coastal waters.
There have been many studies of heavy metals in Australia's marine waters over recent years and in reviewing these data it is important to recognise the problems associated with the sampling and analysis of waters for metals at natural concentrations (Batley 1989). In many instances the significance of the data is questionable, usually because inadequate precautions were taken to avoid contamination during sampling.
Recent studies (Ahlers et al. 1990; Hunter & Tyler 1987) have reinforced the need for extreme care in the sampling of marine waters for trace metal analysis. The earliest concerns were raised by Patterson & Settle (1976) and Schaule & Patterson (1981). More recent analyses of open ocean water samples in which special precautions were taken in preparation of sample containers, sampling operation, and sample handling and analysis, have shown metal concentrations to be lower than originally anticipated. For example, open ocean concentrations of zinc, lead, copper, cadmium and chromium are now estimated to be in the ranges 0.003-0.6, 0.001-0.04, 0.03-0.4, 0.0001-0.12 and 0.1-0.3 µg L-1 respectively ( Bruland et al. 1991; Bruland & Franks 1983; Bruland 1980).
Because of these low concentrations and the problems of contamination, sample handling and analysis must be carried out in a clean laboratory, especially designed so that airborne particles are absent. Unless these facilities have been used, the results of analyses can be considered of questionable validity.
Few analytical techniques are capable of detecting heavy metals at sub microgram per litre concentrations, principally because of interference due to matrix components. For a limited number of metals, anodic stripping voltammetry (ASV) permits direct analysis by in situ electrochemical pre concentration at a mercury electrode, followed by a voltage scan which sequentially dissolves or strips the reduced metals from the electrode, producing a current peak for each which is concentration dependent. Graphite furnace atomic absorption spectrometry (AAS) is the other commonly used technique. For most metals, preliminary extraction with a complexing agent is required to obtain sufficient pre concentration to enable detection of ambient concentrations. Rosman et al. (1982) and Rosman & Dehaeter (1980) have successfully applied isotope dilution mass spectrometry to the detection of cadmium and other heavy metals at nanogram per litre concentrations in waters off Western Australia.
More recently, integrating samplers have been used to preconcentrate heavy metals. These pump large volumes of water (typically >10 L) through a column containing a chelating resin. The metals are eluted with acid and detected by a spectrometric technique such as graphite furnace AAS or inductively coupled plasma mass spectrometry (ICPMS). The method does not necessarily recover total metals, because not all species are dissociated on the resin, and incomplete recovery from the columns is a problem.
Butyltins can be extracted from sediments, waters and biota using non-aqueous solvents preferably in the presence of a complexing agent, tropolone. Reaction with sodium borohydride forms volatile hydrides which can be trapped on a chromatographic column, and separated by thermal desorption, with subsequent detection by quartz furnace atomic absorption spectrometry (Batley et al. 1989b). Alternatively, the hydrides can be solvent extracted and determined directly by gas chromatography using electron capture or flame photometric detection (Batley et al. 1989a). CSIRO's Centre for Advanced Analytical Chemistry in Sydney is the only laboratory in Australia currently experienced in the analysis of TBT in waters at baseline concentrations.
There are few Australian laboratories with the demonstrated ability to determine ultratrace metal concentrations in seawater. Consequently there are few reliable data available for heavy metals in Australian coastal waters. Sampling and analysis of open ocean waters is particularly demanding and is the special expertise of the CSIRO Division of Oceanography in Hobart.
Table 1: Dissolved heavy metal concentrations (µg L-1) in some Australian estuarine waters
|ANZECC Guidelines||5||5||2||50||0.1||50||-||15||50||ANZECC (1992)|
|Central Port Phillip Bay, Vic.||0.6||<0.8||<0.05||<2||<0.002||2.8||-||-||-||EPA (1991)|
|Corio Bay, Vic.||1.1||<0.8||0.2||<2||<0.002||3.2||-||-||-||EPA (1991)|
|Port Hacking Estuary, NSW||0.5||0.4||0.2||-||-||-||-||-||-||Batley and Gardner (1978)|
|North Lake Macquarie, NSW||1.5||1.6||1.9||5.2||-||-||-||-||-||Batley (1987)|
|South Lake Macquarie, NSW||1.2||0.1||0.2||1.0||-||-||-||-||-||Batley (1987)|
|Lake Munmorah, NSW||1.5||0.2||0.1||2.7||0.02||1.7||-||-||2.1||Batley and Brockbank, unpubl.|
|Port Augusta, S.A.||0.45||0.54||0.37||<10||-||-||0||-||-||Ferguson (1983)|
|Port Pirie, S.A. (offshore)||0.25||5.1||0.32||47||-||-||-||-||-||Ferguson (1983)|
|Macquarie Harbour, Tas.a||7||-||0.03||2.0||-||-||-||0.5||-||Carpenter et al. (1991)|
|Derwent River, Tas.||1.2
|Ross River, Qld||0.24
|0.3||Jones and Thomas (1988); Jones (1981)|
a Data are the highest found. Numbers in parentheses are total metals
Metal concentrations are expected to be higher in estuarine waters, and these have naturally been the focus of local and state agencies. Many of the data for heavy metals in estuarine waters have been documented in internal reports of the monitoring agencies, and the resources of this investigation did not enable a detailed survey to be undertaken. Instead, a compilation of results for a number of key estuaries has been reported (Table 1) and evaluated in terms of recently formulated water quality standards (ANZECC 1992). Organisations who use the ANZECC water quality guidelines should also be familiar with the national and site-specific guidelines produced by the USEPA (Stephan et al. 1985). These guidelines provide insights into data base requirements for criteria derivation and relevant laboratory and field data, as well as the toxicological and ecological significance of a substance. In addition USEPA discuss strengths and limitations of criteria and provide guidance on their implication.
The data for estuaries in New South Wales, Victoria, South Australia, Tasmania and Queensland (Table 1) show that in most instances heavy metal concentrations in waters were below the recommended ANZECC guideline concentrations. In some cases measured concentrations are an order of magnitude above seawater concentrations. Notable examples of the latter are Port Pirie (SA), where the impact of effluents from the nearby lead-zinc smelter is clearly seen (Ferguson 1983), and the Lake Macquarie (NSW) sites.
When making comparisons it is important to note that these data are for dissolved metals only, i.e. filterable through a 0.45 µm membrane filter. Total metals, especially for estuarine waters having a high particulate load, can be appreciably higher, as shown in the examples from the Derwent River in Tasmania and the Ross Estuary in Queensland. Considerable emphasis has been given to the determination of the chemical forms (speciation) of metals in waters (Batley 1989; Florence & Batley 1980), since in assessing biological impact it is important to recognise that not all forms of dissolved metals are bio-available. In coastal waters this is generally of little consequence since the concentrations are well below recommended water quality standards.
The trace metal data reported in Tables 1 and 2, are, in many instances, obtained from single samplings. However, concentrations can be quite variable in estuarine and coastal waters, both as a result of changing inputs, or from the temporal effects of biological, chemical and physical interactions. For example, the presence of the blue-green alga Trichodesmium in the Ross River estuary in late winter and spring when the estuary was hypersaline, was accompanied by an increase in copper, lead and cadmium (Jones 1992; Jones & Thomas 1988). At close inshore locations, available copper concentrations were found to increase some 600% to 0.9 µg L-1 during senescence of this alga (Jones et al. 1982), and in later investigations at an offshore reef, similar effects were noted for other trace elements (Jones et al. 1986). The combination of a local source of copper and copper-binding organics released by the algal ligands were believed to be responsible for metal concentrations approaching guideline values.
Table 2: Dissolved heavy metal concentrations (µg L-1) in some Australian coastal waters
|NE Pacific Ocean||0.03||-||-||0.006||-||-||-||-||0.12||-||Bruland et al. (1991)|
|Bate Bay, NSW||0.3||0.2||0.06||-||-||-||-||-||-||-||Batley and Gardner (1978)|
|Pacific Ocean, 8 km off Port Jackson, NSW||<0.2||0.04||0.01||0.1||0.02||1.4||0.08||-||-||0.1||Batley and Brockbank, unpubl.|
|Pacific Ocean off Maroubra, NSW||0.09||0.03||0.01||0.2||0.01||1.0||<0.01||0.04||0.2||0.3||Batley and Brockbank, unpubl.|
|Pacific Ocean off Eden, NSW||0.03||<0.01||<0.005||<0.04||<0.0015||1.5||<0.073||-||0.18||0.1||Apte et al. (1994)|
|Lizard Island, GBR, Qld||0.13||<0.06||<0.01||0.1||<0.002||-||-||-||-||-||Denton and Burdon-Jones (1986)|
|Heron Island, GBR, Qld||0.15||<0.06||<0.01||0.12||<0.001||-||-||-||-||-||Denton and Burdon-Jones (1986)|
|Tasman Sea||0.06||-||0.02||0.08||-||-||-||-||-||0.12||Mackey, unpubl.|
|Indian Ocean, 20 km west of Fremantle||-||0.018||0.002||0.030||-||-||-||-||-||-||Rosman et al. (1982)|
|Cleveland Bay, GBR, Qld||0.1||0.15||0.07||1.02||-||-||-||-||-||0.13||Jones (1981)|
a Total metal concentrations; typical values for surface waters GBR-Great Barrier Reef
Table 3: Trace metals (µg g-1) in some Australian estuarine sediments
|Corio Bay offshore||Surface||4-400||2-210||2-50||0.1-13||Smith (1978)|
|Corio Bay||Surface||14-166||14-100||4-35||0.2-9||Fabris (1983)|
|Port Phillip Bay, near shore||Surface||21||8||1.5||0.8||Talbot et al. (1976)|
|Port Phillip Bay, offshore||Surface||40||22||8||2||EPA (1976)|
|Port Phillip Bay, near Werribee Treatment Complex||Surface||9-300||<20-140||<5-75||<5||MBW (1991)|
|-||Batley et al. (1990)|
|-||Batley et al. (1990)|
|Lake Macquarie North||55||2400||1200||170||160||Batley (1987)|
|Lake Macquarie South||150||68||20||4|
|Blackwattle Bay, Sydney Harbour||10||1150||520||180||3||Batley (1986)|
|Port Kembla Harbour||10||380||113||113||2||Batley and Low (1985)|
|Quibray Bay, Botany Bay||10||25||10||3||0.5||Batley, unpubl.|
|Sydney coast (100m water depth), typical high values for clay-silt||Surface||60||15||14||-||Batley and Brockbank, unpubl.|
|Central GBR (John Brewer Reef)||Surface||5||0.6||0.2||-||Jones (1992)|
|Cleveland Bay, Qld||10||24-460||<5-53||15-70||-||Reichelt and Jones (1992)|
|Bowling Green Bay||Surface||10-26||<0.5-4||1-4||-||Burdon-Jones et al. (1977|
Metal concentrations in coastal waters are usually much lower than estuarine waters (Table 2), with values for both well below those reported for northern hemisphere coastal waters where industrialisation has had a greater impact, and where discharge standards have been more relaxed (Forstner & Wittman 1979). Values for dissolved metals did not differ greatly from waters of the Great Barrier Reef to those in the Pacific Ocean off New South Wales or the Indian Ocean off Western Australia. The detection limits for these data were lowest in the latter samples where isotope dilution mass spectrometry was used as the method of analysis. Concentrations were higher closer to shore, for example copper in Bate Bay (NSW) or dissolved zinc in Cleveland Bay Qld), suggesting local contamination. A recent detailed study by Apte et al. (1994) of waters from sites off the NSW coast, using ultraclean techniques for both sampling and analysis, has yielded data which are amongst the lowest reported for Australian waters.
In estuarine and inshore waters, as a result of both particulate and dissolved inputs, heavy metals will be enriched in suspended sediments as they are in the dissolved phase, and there is now substantial evidence of heavy metal enrichment in the bottom sediments of coastal waters. This is greatest in sediments close to the mouth of estuaries, reflecting their role as a contaminant source. Enrichment will be a function of grain size, being greatest with the smaller area, higher surface area clay and silt particles than with sandy sediments.
Data for some Australian sediments are shown in Table 3. The lowest numbers reflect a high sand content in the samples. Industrial and port activities, stormwater run-off and sewage discharges have been identified, in the studies quoted, as the major sources of the high concentrations observed for metals such as zinc, lead and copper.
Speciation of metals in sediments is sometimes necessary to delineate pollution sources, especially to distinguish between mineralised or lattice-held metals and the more bio-available fractions (Kersten & Forstner 1989). For this purpose, the use of dilute acid or complexing agents such as ethylenediamine tetraacetic acid (EDTA) as extractants has been found to yield concentrations which most closely relate to the bio-available fraction, and this fraction has been measured in sediments from a number of Australian estuaries (Reichelt and Jones 1992; Batley, 1987).
The definition of sediment quality guidelines is still the subject of extensive research both in Australia and abroad. While no definitive guidelines have yet been published, typically concentrations of 100 µg Zn g-1, 50 µg Pb g-1, 25 µg Cu g-1 and 8 µg As g-1 have been considered as guidelines for dredged sediment disposal in Canada, and these are not too dissimilar to suggested EPA values. On this basis, most of clay-silt sediments in urban and industrialised estuaries around Australia would exceed the guideline values (Table 3).
Table 4: Zinc and copper ( µg g-1) in some Australian oysters
|Hook Island, close to resort, Great Barrier Reef||1225-9648
|Hook Island, distant from resort, Great Barrier Reef||131-1314
|Rattlesnake Island, Great Barrier Reefb||1441-2095
|Burdon-Jones et al. (1977)|
|Dampier Archipelago, WA||55-1800a||31-200a||Talbot (1985)|
|Shark Bay, Vic.||(180)a||-||McConchie et al. (1988)|
|Darwin Harbour, NT||109-611a||18-58a||Peerzada and Dickinson (1988)|
|Townsville Harbour, Qldb||673-20 906
|Jones (1981), Jones (1992)|
|Georges River, NSW||80-665a||3-48a||Mackay et al. (1975)|
|Georges River, NSW||440-760a
|Batley et al. (1992)|
|Tasman River, Tas.||1700-14 000||200-1700||Ayling (1974)|
a Wet weight.
b Seasonal study.
Numbers in parentheses are total metals
There is a plethora of data for metals in biota because of the importance of food chain bio-accumulation and ultimate human consumption. The National Health and Medical Research Council (NH&MRC) set guideline concentrations for heavy metals in fish, crustaceans and molluscs. There have been isolated instances of certain organisms exceeding these guideline concentrations and this has generally been attributable to point source pollution from urban or industrial sources. Such contamination has often been detected in waters and sediments from where the biota was collected, indicating the route of metal uptake. More often, because organisms bio-accumulate and effectively integrate metal loads, biota provide a more reliable measure of the presence of metal pollution in waters. It should be noted that bio-accumulation will be dependent on the chemical form of heavy metals, with ionic species being generally more bio-available than bound or complexed forms. The analysis of biota is also a less exacting task than that of waters. There are a number of examples where uptake and biomagnification of metal concentrations occurs via the food chain, for example through algae or aquatic plants, but the original source is always water or sediment.
Mercury is a metal that has received considerable attention since the early 1970s. Recent measurements of fish species, at the top of the food chain, still show tissue concentrations exceeding the NH&MRC guideline value of 0.5 µg g-1 (wet wt), at sites such as Port Phillip Bay (Walker 1982, 1981) and Townsville's coastal waters (Denton & Breck 1981). Elevated concentrations of other metals have in the past been detected in fish, oysters and seagrasses near smelters, refineries and other heavy industry. The number of such instances has now been reduced in New South Wales and Victoria because of more stringent discharge controls, and this is likely to occur in other States.
In tropical Queensland waters an extensive investigation of metal levels in a wide range of organisms collected from urbanised and industrial locations has been undertaken (Burdon-Jones et al. 1977; Burdon-Jones et al. 1975). These studies highlighted significant metal contamination in various marine organisms collected from sites close to port activities, urbanisation and the discharge of sewage. One of the objectives of this study was to identify useful monitoring organisms. The oysters, Saccostrea amassa and Saccostrea echinata fulfilled many of the requirements of a useful monitoring organism in that they are abundant in the intertidal zone, sessile, euryhaline, long lived, and exhibit high concentration factors for a number of metals. Mean concentrations of zinc exceeded the NH&MRC guideline values in S. amassa from Cleveland Bay (Table 4) close to port activities and sewage discharges (Jones 1992) and were comparable to values for oysters from the polluted Tamar Estuary in Tasmania (Ayling 1974). The upper range of zinc concentrations (38,700 µg g-1 dry weight) was close to the highest recorded concentration of zinc in oysters (Bloom & Ayling 1977). Copper concentrations were even higher than reported for the Tamar Estuary (Jones 1992). Comparison of zinc concentrations in oysters from a range of locations throughout Australia clearly indicates their usefulness as a monitoring organism (Table 4).
Table 5: Tributyltin in Australian coastal waters
|Site||Date||TBT (ng Sn L-1)||Reference|
|Georges River, NSW||Pre-ban||8-40||Batley et al. (1989a)|
|Georges River, NSW||Post-ban||1-11||Batley and Scammell, unpubl.|
|Kogarah Bay, NSW||Pre-ban||100||Batley et al. (1989a)|
|Port Phillip Bay, Vic.||Pre-ban||3-23||Batley and Scammell (1991)|
|Southport, Qld||Pre-ban||45||Batley and Scammell (1991)|
In the Georges River (NSW), an urban estuary, oysters (Saccostrea commercialis) in 1988 had copper concentrations exceeding the guideline concentration of 70 µg g-1. This was attributed to antifouling paints (Batley et al. 1992). Copper has long been a component of antifouling paints but mixtures of copper and TBT have been found to bee more effective against the range of biofouling organisms. Copper was elevated in oysters because of a synergism involving TBT, but following the banning of the latter in paints, the copper concentrations returned to below the limit.
The guideline concentration for TBT in marine waters is 2 ng L-1. Prior to banning, measurements had been obtained for water samples from a number of sites in NSW and other states (Table 5). Particularly near dockyards or other areas of high shipping densities, concentrations of TBT in excess of 100 ng Sn L-1 were found. In waterways inaccessible to large vessels, concentrations ranged from 10-100 ng Sn L-1. Following banning, concentrations were in most instances near or below the guideline value.
As with heavy metals, the greatest concentrations of TBT are found in sediments. Data have been obtained for sites in Vic., Qld, SA, WA and NSW for their respective environmental authorities, and the results have been similar for each location (Batley & Scammell 1991; Witney 1991; EPAWA 1990).
Typically concentrations in sediments are low and are of minor environmental concern. In the vicinity of marinas, however, sediment concentrations can exceed 1 µg Sn g-1 or higher, the latter usually the result of paint flakes hydroblasted from boat hulls (Batley & Scammell 1991). Surveys of estuaries around Sydney have shown that, following banning, TBT concentrations are now lower in the upper 1-3 cm of sediment, depending on sedimentation rate. A maximum concentration appears below this depth, diminishing in deeper sediments consistent with its successive degradation to dibutyltin, monobutyltin and inorganic tin.
The range of data on TBT in Australian aquatic biota has been summarised in several publications (Batley & Scammell 1991; Maher & Batley 1990; Scammell et al. 1990). The most significant impacts have been on intertidal oysters but with an apparent lack of impact on subtidal oysters. The reason for this is postulated to be related to enrichment of TBT in the surface microlayer. Reductions in populations of scallops in Victoria might also be attributable to the impact of TBT in the microlayer on larval scallops.
Both problems have now been reversed with the banning of TBT (Batley et al. 1992). Oyster growth is now normal, with TBT barely detectable, and larger than ever scallop populations are being reported in Port Phillip Bay.
The observations in gastropods of imposex (the induction of male reproductive organs in female animals) caused by TBT has been examined in a number of sites in NSW (Ahsanullah and Wilson, unpublished results). The impact of banning on this phenomenon has not been reported. There appears to have been no significant impacts or accumulation of TBT by other aquatic biota.
While this review has concentrated principally on the eastern seaboard of Australia, at sites known to the author, the results can be considered typical of the continent's coastal waters. In general heavy metal concentrations in coastal waters are low and in most cases approach open ocean values, although the data base is small. In estuaries, concentrations are higher and in limited cases where point source inputs are responsible, values are found which exceed water quality guidelines. This represents an improvement in conditions of a decade or two earlier, and reflects improved discharge controls for some states. TBT concentrations in estuarine waters have also decreased to close to or below ANZECC guideline values since its banning in antifouling paints for small boats.
The impacts from lowering dissolved metals in waste discharges is seen in recent data for sediments and biota. Surface sediments generally show lower metal concentrations than at depth, but the top 50 cm of most urbanised and industrialised estuaries are contaminated with heavy metals, especially lead and zinc. Biota accumulating metals from either waters or sediments, can reflect significant contamination, and in some cases this leads to values in excess of the NH&MRC guidelines.
In conclusion, however, it is felt that comparable baseline studies are still needed for other coastal zones and estuaries throughout Australia. Furthermore, knowledge of the physical and chemical processes in coastal waters is essential in order to understand the mechanisms involved in the distribution of pollutants from their points of entry. Care needs to be taken when extrapolating water quality guidelines to tropical environments where considerable variation in temperature and salinity can occur at different times of the year.
Acknowledgments: The assistance of Dr Denis Mackey, CSIRO Division of Oceanography, and Dr Graham Jones, James Cook University of North Queensland in reviewing this chapter and providing additional data, is gratefully acknowledged.
Ahlers, W.W., Reid, M.R., Kim, J.P., & Hunter, K.A. 1990, 'Contamination-free sample collection and handling protocols for trace elements in natural fresh waters', Australian Journal of Marine and Freshwater Research, vol. 41, pp. 713-720.
ANZECC 1992, Australian Water Quality Guidelines for Fresh and Marine Waters, Australian and New Zealand Environment and Conservation Council.
Apte, S.C., Batley, G.E., Szymczac, R., Waite, T.D., Rendell, P.S., & Lee, R. 1994, 'Baseline trace metal concentrations in New South Wales coastal waters', Australian Journal of Marine and Freshwater Research, submitted.
Ayling, G.M. 1974, 'Uptake of cadmium, zinc, copper, lead and chromium in the Pacific oyster, Crassostrea gigas, grown in the Tamar River, Tasmania', Water Research, vol. 8, pp. 729-738.
Batley, G.E. 1989, 'Collection, preparation, and storage of samples for speciation analysis', in Trace Element Speciation, Analytical Methods and Problems, ed G.E. Batley, CRC Press, Boca Raton, pp. 1-24.
Batley, G.E., Body, N., Boon, P., Cook, B., Dibb, L., Fleming, P.M., Mitchell, D., Sinclair, R., & Skyring G. 1990, The ecology of the Tuggerah Lakes System. A review with special reference to the impact of the Munmorah Power Station, Joint Murray Darling Freshwater Research Institute/CSIRO Investigation Report.
Batley, G.E., Brockbank, C.I. & Scammell, M.S. 1992, 'The impact of banning of tributyltin-based antifouling paints on the Sydney rock oyster, Saccostrea commercialis,' Science of the Total Environment, vol. 122, pp. 301-314.
Batley, G.E., Chen Fuhua, Brockbank, C.I. & Flegg, K.J. 1989a, 'Accumulation of tributyltin in the Sydney rock oyster, Saccostrea commercialis,' Australian Journal of Marine and Freshwater Research, vol. 40, pp. 49-57.
Batley, G.E. & Low, G. K-C. 1985, Heavy metals and polycyclic aromatic hydrocarbons in sediments from a dredging site in Port Kembla Harbour, Chemistry Investigation Report EC/IR013, CSIRO Division of Energy .
Batley, G.E. & Scammell, M.S. 1991, 'Research on tributyltin in Australian estuaries', Applied Organometallic Chemistry, vol. 5, pp. 99-105.
Bloom, H. & Ayling, G.M. 1977, 'Heavy metals in the Derwent Estuary', Environmental Geology, vol. 2, pp. 3-22.
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