Cadmium is widely distributed in the Earth's crust at an average concentration of about 0.1 mg kg^-1 and is commonly found in association with zinc. Higher levels are present in sedimentary rocks: marine phosphates often contain about 15 mg kg^-1 (GESAMP, 1984). Weathering and erosion result in the transport by rivers of large quantities of cadmium to the world's oceans and this represents a major flux of the global cadmium cycle; an annual gross input of 15,000 tonnes has been estimated (GESAMP, 1987). Volcanic activity is also a major natural source of atmospheric cadmium release. The global annual flux from this source has been estimated to be 100-500 tonnes (WHO 1992b).

The principal applications of cadmium fall into five categories:

• protective plating on steel;
• stabilizers for PVC;
• pigments in plastics and glass;
• electrode material in nickel-cadmium batteries; and
• as a component of various alloys (Wilson, 1988).

The relative importance of the major applications has changed considerably over the last 25 years. The use of cadmium for electroplating represented has decreased, with its share in 1985 less than 25% (Wilson, 1988). In contrast, the use of cadmium in batteries has grown considerably in recent years from only 8% of the total market in 1970 to 37% by 1985.

Non-ferrous metal mines represent a major source of cadmium to the aquatic environment. Contamination can arise from mine drainage water, waste water from the processing of ores, overflow from the tailings pond, and rainwater run-off from the general mine area. The release of these effluents to local watercourses can lead to extensive contamination downstream of the mining operation. Disused mines can also be a source of water contamination (Johnson and Eaton, 1980).

At the global level, the smelting of non-ferrous metal ores has been estimated to be the largest human source of cadmium release to the aquatic environment (Nriagu and Pacyna, 1988). Discharges to fresh and coastal waters arise from liquid effluents produced by air pollution control (gas scrubbing), together with the site drainage waters.

The manufacture of phosphate fertilizer results in a redistribution of the cadmium present in the rock phosphates between the phosphoric acid product and gypsum waste. In many cases, the gypsum is disposed of by dumping in coastal waters, which leads to considerable cadmium inputs. The atmospheric fall-out of cadmium to fresh and marine waters also represents a major input of cadmium at the global level (Nriagu and Pacyna, 1988). A GESAMP study of the Mediterranean Sea indicated that this source was comparable in magnitude to the total river inputs of cadmium to the region (GESAMP, 1985). Similarly, large cadmium inputs to the North Sea (110-430 tonnes/year) have also been estimated, based on the extrapolation from measurements of cadmium deposition along the coast (van Alst et al., 1983a,b). However, another approach based on model simulation yielded a modest annual cadmium input of 14 tonnes (Krell and Roeckner, 1988).

The average cadmium content of sea water has been given as about 0.1 &micro;g l^-1 or less (Korte, 1983). WHO (1992) reported that current measurements of dissolved cadmium in surface waters of the open oceans gave values of < 5 ng l^-1. The vertical distribution of dissolved cadmium in ocean waters is characterized by a surface depletion and deep water enrichment, which corresponds to the pattern of nutrient concentrations in these areas (Boyle et al., 1976). This distribution is considered to result from the absorption of cadmium by phytoplankton in surface waters and its transport to the depths, incorporation to biological debris, and subsequent release. In contrast, cadmium is enriched in the surface waters of areas of upwelling and this also leads to elevated levels in plankton unconnected with human activity (Martin and Broenkow, 1975; Boyle et al, 1976). Oceanic sediments underlying these areas of high productivity can contain markedly elevated cadmium levels as a result of inputs associated with biological debris (Simpson, 1981). Cadmium levels of up to 5 mg kg^-1 have been reported in river and lake sediments and from 0.03 to 1 mg kg^-1 in marine sediments (Korte 1983).

Concentrations of cadmium have been measured in water, sediments and biota as part of the National Monitoring Programme at sites throughout the UK in estuaries and coastal waters (MPMMG 1998). The results of the National Monitoring Programme are summarised in Appendix D. MPMMG 1998 should consulted for further details.

The available data suggest that, although sometimes elevated, in general, concentrations of cadmium in UK coastal and estuarine waters, sediments and biota do not appear to exceed relevant quality standards derived for the protection of saltwater life.

As an example of the recorded levels of dissolved cadmium in the marine environment, the following concentrations have been reported by DETR (1998) for some English estuaries (See tables below).

Cadmium does not break down in the environment, but may be affected by physical and chemical processes that modify its mobility, bioavailability, and residence time in different environmental media. The mobility and bioavailability of cadmium in aquatic environments are enhanced under conditions of low pH, low hardness, low suspended matter levels, high redox potential, and low salinity.

The speciation of cadmium in the environment is of importance in evaluating the potential hazard. Some cadmium salts, such as the sulphide, carbonate, and oxide, are practically insoluble in water; these can be converted to water-soluble salts in nature. The sulphate, nitrate, and halides are soluble in water. In addition, cadmium is strongly adsorbed onto sediments.

An exhaustive literature review on the toxicity of cadmium to marine organisms has not been carried out for the purposes of this profile. The information provided in this section is taken from existing review documents (WHO 1992b). The most sensitive groups of organisms have been identified.

Cadmium uptake from water by aquatic organisms is extremely variable and depends on the species and various environmental conditions, such as water hardness (notably the calcium ion and zinc concentration), salinity, temperature, pH, and organic matter content.

The majority of chelating agents decrease cadmium uptake but some, such as dithiocarbamates and xanthates, increase uptake. Increasing temperature increases the uptake and toxic impact, whereas increasing salinity or water hardness decreases them.

Acute lethal effects for marine organisms have been noted as low as 16 &micro;g l^-1 (WHO 1992b).

Cadmium is toxic to a wide range of micro-organisms. However, the presence of sediment, high concentrations of dissolved salts or organic matter all reduce the toxic impact. The main effect is on growth and replication.

Zinc increases the toxicity of cadmium to aquatic invertebrates. Sub-lethal effects have been reported on the growth and reproduction of aquatic invertebrates; there are structural effects on invertebrate gills. There is evidence of the selection of resistant strains of aquatic invertebrates after exposure to cadmium in the field.

An increase in toxicity as temperature increases and salinity decreases has been noted. This implies that the same cadmium concentration may have the potential to cause greater toxicity to estuarine rather than to marine species. For example, Rosenberg and Costlow (1976) reported increased cadmium toxicity during larval development of two estuarine crab species as salinity decreased and increased toxicity as temperature increased.

O'Hara (1973) investigated the effects of temperature and salinity on the toxicity of cadmium to adult male and female fiddler crabs Uca pugilator. Mortality was greatest at high temperatures and low salinities in tests lasting 240 h. LC50 values varied from 2.9 mg l^-1 for the lowest salinity (10%) and highest temperature (30 °C) to 47.0 mg l^-1 for the highest salinity (30%) and lowest temperature (10 °C). Frank and Robertson (1979) exposed the blue crab Callinectes sapidus to cadmium chloride at salinities of 1, 15, and 35%. Like O'Hara, they found a decrease in cadmium toxicity with increase in salinity. For example, 96-h LC50 values were 0.32, 4.7, and 11.6 mg cadmium l^-1 for the three salinities respectively.

Voyer and Modica (1990) found the same pattern with the mysid shrimp Mysidopsis bahia. For salinities of 10 and 30%, the 96-h LC50 values ranged from 15.5 to 28 &micro;g cadmium l^-1 at a temperature of 25 °C and from 47 to 84 &micro;g l^-1 at a temperature of 20 °C. At 30 °C, the 96-h LC50 was < 11 &micro;g l^-1 at both salinities. However, when Robert and His (1985) exposed embryos and larvae of the Japanese oyster Crassostrea gigas to cadmium concentrations of up to 50 &micro;g l^-1 at various salinities (20 to 35%), decreasing the salinity severely affected the development of the oysters but cadmium had no effect. At temperatures higher than 11 °C, the combined effect of temperature and cadmium caused a heavy stress to the copepod Tisbe holothuriae so that the effects of salinity were masked (Verriopoulos and Moraitou-Apostolopoulou, 1981).

At low concentrations (10 &micro;g cadmium l^-1), cadmium inhibits ion transport systems and induces metallothionein synthesis (< 1 &micro;g cadmium l^-1) in freshwater fish.

Cadmium toxicity has been found to be variable in fish, with salmonids being particularly susceptible to cadmium. Sub-lethal effects in fish, notably malformation of the spine, have been reported. The most susceptible life-stages are the embryo and early larva, while eggs are the least susceptible. There is no consistent interaction between cadmium and zinc in fish (WHO 1992b).

Cadmium has been found to affect the kidney of sea-birds. However, it is not always thought to have been as a result of exposure to cadmium as an industrial pollutant since the individuals most affected come from areas where there is no industrial effluent. Such effects appear therefore to be a response to naturally occurring cadmium presumed to derive from the oceans (WHO 1992b).

However, these effects do not appear to affect survival or breeding success. No damage resulting from exposure to strictly anthropogenically derived cadmium appears to have been reported on the same scale as that from exposure to naturally occurring cadmium (WHO 1992b).

Nicholson et al (1983) compared the kidneys of seabirds contaminated with cadmium in the wild, seabirds from uncontaminated colonies, starlings dosed with cadmium in the laboratory, and control starlings. The authors found damage to kidney cells to be comparable between wild seabirds and dosed starlings having kidney cadmium levels of 60-480 and 95-240 mg kg^-1 respectively. Nicholson and Osborn (1983) reported kidney lesions in several different species of seabird caught in contaminated areas, although other pollutant metals, such as mercury, were also present in the tissues.

Cadmium bioaccumulates in organisms with the main uptake routes being dissolved cadmium from the water column and cadmium associated with prey items.

Muller et al (1993) reported bioconcentration factors of 5,000 in natural samples of Enteromorpha intestinalis in the Weser estuary and these increased to between 7,000 and 10,000 for plants transplanted from a clean source.

The potential effects include:

• acute toxicity of dissolved cadmium to invertebrates and fish at concentrations above the EQS of 2.5 &micro;g l^-1 (annual average) in the water column;
• accumulation of cadmium in sediments and a potential risk to sediment dwelling organisms at concentrations greater than 0.7 mg kg^-1, according to Canadian interim marine sediment quality guidelines;
• bioaccumulation of cadmium in organisms and a potential threat to some invertebrates, fish, birds and Annex II mammals.

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