Copper in freshwater and marine water

​​Toxicant default guideline values for protecting aquatic ecosystems

October 2000

Extracted from Section 8.3.7 ‘Detailed descriptions of chemicals' of the ANZECC & ARMCANZ (2000) guidelines.

The default guideline values (previously known as ‘trigger values’) and associated information in this technical brief should be used in accordance with the detailed guidance provided in the Australian and New Zealand Guidelines for Fresh and Marine Water Quality.

Description of chemical

Copper is found at low concentrations in most marine, estuarine and fresh waters (Table 8.3.2 of the ANZECC & ARMCANZ 2000 guidelines). Copper is an essential trace element required by most aquatic organisms but toxic concentrations are not much higher than those that allow optimum growth of algae. Cairns et al. (1978) noted that copper stimulated growth of Scenedesmus quadricauda and Chlamydomonas sp. at near lethal concentration. It is generally assumed that the free hydrated copper ion (Cu2+) together with copper hydroxy species are the most toxic inorganic species to aquatic organisms.

Copper is readily accumulated by plants and animals; bioconcentration factors ranging from 100 to 26,000 have been recorded for various species of phytoplankton, zooplankton, macrophytes, macroinvertebrates and fish (Spear & Pierce 1979). Toxic effects of metals occur when the rate of uptake exceeds the rates of physiological or biochemical detoxification and excretion (Rainbow 1996). This is more important than absolute body burden. Jarvinen and Ankley (1999) report data on tissue residues and effects for copper for 14 freshwater species and 9 marine species. It is not possible to summarise the data here but readers are referred to that publication for more information. Ahsanullah and Williams (1991) reported that the marine amphipod Allorchestes compressa exposed to 10 µg/L of copper for 28 days accumulated 100 mg/kg of copper and experienced reduced growth.

Summary of factors affecting copper toxicity

  • Copper is an essential trace element required by many aquatic organisms.
  • Copper toxicity decreases with increasing hardness and alkalinity and a hardness algorithm is available (Table 3.4.3 of the ANZECC & ARMCANZ 2000 guidelines).
  • Levels of dissolved organic matter found in most freshwaters are generally sufficient to remove copper toxicity but often not in very soft waters. Speciation measurements can account for this.
  • Copper is adsorbed strongly by suspended material. Filtration and speciation measurements should account for this.
  • Copper complexing is increased at higher pH, but the relationship to toxicity is complex.
  • Copper toxicity in algae, invertebrates and fish generally increases as salinity decreases.
  • Copper can bioaccumulate in aquatic organisms but, as it is an essential element, it is commonly regulated by the organisms.

A variety of methods are available for determining the speciation of copper in water including:

  1. Analytical techniques, such as physical separation (e.g. (ultra)filtration, dialysis, centrifugation), potentiometry (e.g. ion-selective electrode), polarography, voltammetry (e.g. anodic stripping voltammetry), ion exchange and ligand competition methods (Florence & Batley 1980, Harrison & Bishop 1984, Apte & Batley 1995, Xue & Sunda 1997)
  2. Theoretical techniques, such as geochemical modelling (Sylva 1976, Leckie & Davis 1979, Miwa et al. 1989).

Bioassays are typically used to ascertain metal-organism interaction. These can be coupled with the measured and/or predicted speciation of copper to determine the bioavailable copper species. The current analytical practical quantitation limit (PQL) for copper is 0.1 µg/L in freshwater and 2 µg/L in marine water (NSW EPA 2000).

Factors that affect the toxicity of copper

In natural waters, copper is largely complexed by natural dissolved organic matter (DOM) such as humic, fulvic and tannic acids, or adsorbed to colloidal, humic-coated iron and/or manganese oxide particles (Mantoura et al. 1978, Florence & Batley 1980, Moore & Ramamoorthy 1984b). Most copper in natural waters is present as copper-DOM complexes. The more toxic inorganic copper species comprise only a relatively minor proportion of the dissolved copper pool in coastal waters, estuaries and rivers (van den Berg 1984, van den Berg et al. 1986, 1987, Apte et al. 1990, Xue & Sigg 1993).

The complexation of copper with DOM increases in freshwaters as the pH and concentration of DOM are increased, and as the concentrations of competing ions are decreased (Sylva 1976). In most natural waters, the concentration of available dissolved organic complexing ligands (copper complexation capacity) is greatly in excess of total dissolved copper and this ensures that inorganic copper concentrations are well below concentrations of toxicological concern. In freshwaters, particularly acidic soft waters with low complexation capacity, copper may be highly toxic. The copper complexation capacity of the surface waters of the Northeast Pacific Ocean is typically 0.6 µg/L, whereas dissolved copper concentrations are less than 0.05 µg/L (Coale & Bruland 1988). In estuaries, the copper complexation capacity ranges from 0.9 to 35.0 µg/L (van den Berg et al. 1986, 1987, Apte et al. 1990, Gardner& Ravenscroft 1991a,b).

The vast majority of studies have shown that natural dissolve organic matter (e.g. fulvic and humic acids) reduce the uptake and toxicity of copper in freshwater organisms (e.g. Meador 1991,Welsh et al. 1993, Azenha et al. 1995, Ericksonet al. 1998a, Hanstén et al. 1996). However, several studies have shown that some of these organic complexing agents may enhance the uptake and toxicity of copper under certain conditions (Guy & Kean 1980, Daly et al. 1990, Tubbing et al. 1994, Buchwalter et al. 1996).

Sorption onto minerals, clays and biotic surfaces and precipitation play major roles in determining the fate of copper (II) in aquatic systems (Dzombak & Morel 1990, Goldberg et al. 1996, Tessier et al. 1996). Sorption of copper (II) to oxyhydroxides increases with pH, until a threshold point is reached usually around pH 8 (Dzombak & Morel 1990).

The uptake and toxicity of copper in freshwater organisms generally decreases with increasing water hardness and alkalinity (Erickson et al. 1998a) [also see reviews by Sorenson (1991) and Mayer et al. (1994b)]. For example, Gauss et al. (1985) reported that the 96-h EC50 for a chironomid (Chironomus tentans) was 17 µg Cu/L in soft water (hardness, 43 mg/L as CaCO3; alkalinity, 32 mg/L as CaCO3; pH 7.6). In contrast, it was 98 µg Cu/L in hard water (hardness, 172 mg/L as CaCO3; alkalinity, 111 mg/L as CaCO3; pH 8.1). An exponential, inverse relationship has been shown to exist between water hardness and the uptake and toxicity of copper. An algorithm describing this relationship has been used to calculate a hardness-modified copper guideline value for protecting aquatic ecosystems in North America (USEPA 1995a,b).

Conflicting results have been reported on the effect of pH (H+) on the uptake and toxicity of copper in freshwater organisms. Most studies have reported that the uptake and toxicity of copper decreases with decreasing pH (e.g. Campbell & Stokes 1985, Cusimano et al. 1986, Macfie et al. 1994,Horne & Dunson 1995, Erickson et al. 1998a) over the pH range 3 to 7. However, some studies have shown an increase in the uptake and toxicity of copper with decreasing pH (Waiwood & Beamish 1978, Schubauer-Berigan et al. 1993), over the pH range 6.0 to 8.5.

Partitioning and bioaccumulation of copper in natural waters is controlled by active biological processes as much as by chemical equilibria. Organisms such as algae and fish release dissolved organic ligands, which bind copper and control its uptake and bioavailability. Exudate production is dependent on the copper concentration, nutrients and physiological status of the organisms (Zhou et al. 1989). Erickson et al. (1998a) recognised that toxicity is also affected by water chemistry in ways not related to copper speciation.

Copper toxicity in algae, invertebrates and fish generally increases as salinity decreases (Denton & Burdon-Jones 1982, 1986, Stauber 1995, Stauber et al. 1996b). For example, the toxicity of copper to juvenile banana prawns Penaeus merguiensis increased with decreasing salinity (Denton & Burdon-Jones 1982). The 96-hour LC50 decreased from 6.1 mg/L at 36‰ salinity to 0.72 mg/L at 20‰ salinity. Copper toxicity to banana prawns also increased with increasing temperature.

Aquatic toxicology

USEPA (1985a) reported acute toxicity data for copper in freshwater species in 41 genera. At a hardness of 50 mg/L, the values ranged from 17 µg/L for Ptychocheilus to 10,000 µg/L for Acroneuria. Skidmore & Firth (1983) found the acute toxicity of copper for ten Australian species ranged from 200 µg/L to 7800 µg/L. Bacher & O’Brien (1990) reported a range for Australian species ranged from 40 µg/L to 21,000 µg/L.

Fish and invertebrates seem to be about equally sensitive to the chronic toxicity of copper in fresh waters. The sensitivity of a number of species of freshwater plants that were tested was similar to those of animals (USEPA 1986). Copper and lead appeared to interact with synergism, both with sequential and simultaneous exposure (Tao et al. 1999).

Some species of algae are particularly sensitive to copper and both marine and freshwater algae vary considerably in their sensitivity. The concentrations of copper reported to cause a 50% decrease in algal growth ranged from 5 to 58,000 µg/L (Fisher et al. 1981, Gavis et al. 1981). The large range could be explained by the use of culture media that contain chelators and absorbents, which reduce copper toxicity. Using unsupplemented seawater, or synthetic soft water (hardness 30 to 40 mg CaCO3​/L) enriched only in nitrate and phosphate, Stauber and Florence (1989) found 72-hour EC50 values of 10 µg/L and 16 µg/L of Cu for Australian isolates of a marine and freshwater alga respectively. Toxicity of copper to the freshwater alga decreased with increasing water hardness.

The acute toxicity of copper to saltwater animals ranged from 5.8 µg/L for blue mullet to 600 µg/L for green crab (USEPA 1986). Invertebrates, particularly marine crustaceans, corals and sea anenomes, are sensitive to copper, with concentrations of copper as low as 10 µg/L causing sublethal effects. Acute LC50 values for prawns, crabs and amphipods ranged from 100 to 1000 µg/L, with chronic values from 10 to 300 µg/L (Arnott & Ahsanullah 1979, Ahsanullah & Florence 1984).

Gastropods are more tolerant to copper and can accumulate quite high concentrations without toxic effects. Typical 96-h LC50 values for snails are 0.8 to 1.2 mg Cu/L. Marine bivalves, including the mussel Mytilus edulis are more sensitive to copper, with a 96-hour LC50 of 480 µg/L (Amiard-Triquet et al. 1986). Growth reductions were found at copper concentrations as low as 3 µg Cu/L. Larvae of the doughboy scallop Mimachlamys asperrimus and the Pacific oyster Crassostrea gigas are among the most sensitive Australian species. Larval development of these is inhibited at copper concentrations as low as 3 µg/L. Toxic effects on saltwater algae were observed at copper concentrations between 5 µg/L and 100 µg/L (USEPA 1985a).

Marine fish appear to be relatively tolerant of copper. The 96-hour LC50 for Australian juvenile glass perch and diamond-scaled mullet were reported as 2000 to 6000 µg Cu/L (Denton & Burdon-Jones 1986). In general, embryos of marine fish are more sensitive than their larvae whereas larvae of freshwater fish are more sensitive than embryos (Rice et al. 1980). Some freshwater fish species, especially salmonoids, are more sensitive to copper than marine fish, with 96-hour LC50 values 40 to 80 µg/L in soft waters and 250 µg/L in hard waters. Freshwater carp are more resistant to copper, with 96-hour LC50 values typically 250 µg/L in soft waters and 3000 µg/L in hard waters.

Freshwater guidelines

For freshwater guideline derivation, only the chronic data that were linked to pH and hardness measurements were considered and further screened. This reduced the dataset to around 130 data points covering 4 taxonomic groups, and these were adjusted to a common hardness of 30 mg/L as CaCO3, as follows (data are reported as geometric means of NOEC after adjustment from other chronic end-points (van de Plassche et al. 1993) (pH range was 6.96 to 8.61):

Fish: 10 species, 2.6 µg/L (Ptylocheilus oregonensis, from 7-day LC50) to 131 µg/L (Pimephales promelas, 7-day LC50); seven species had geometric means <25 µg/L

Crustaceans: five species, 1.7 µg/L (D. pulex and G. pulex, NOEC, reproduction & mortality) to 12.1 µg/L (Hyalella azteca, from 10 to 14-day LC50)

Insects: three species, 2.2 µg/L (Tanytarsus dissimilis, from 10-day LC50) to 11 µg/L (Chironomus tentans, 10 to 20-day LC50)

Molluscs: three species, 1.64 µg/L (Flumicola virens, from 14-day LC50) to 56.2 (Corbicula manilensis, from 7 to 42-day LC50). The latter figure was not included in calculations as it was outside the pH range.

A freshwater high reliability trigger value for copper of 1.4 µg/L was derived using the statistical distribution method with 95% protection. This applies to waters of hardness of 30 mg/L as CaCO3.

Marine guidelines

Screened data consisted of 70 data points from five taxonomic groups, as follows (expressed as geometric means of NOEC equivalents; pH data were not recorded):

Fish: six species, 30 mg/L (two species, from 12 to 14-day EC50, hatch & mortality) to 260 µg/L (Menidia menidia, 11-day EC50, hatch).

Crustaceans: three species, 1.7 µg/L (Callianassa australiensis, from 10 to 14-day EC50 of 8.5 mg/L) to 42 µg/L (Mysidiopsis bahia, from 29 to 51-day MATC, reproduction).

Molluscs: seven species, 0.4 µg/L (Mytilus edulis, from 30-d EC50, reproduction of 2 mg/L) to 20 000 µg/L (Ostrea edulis, 5-d LC50).

Annelids: three species, 17 µg/L to 68 µg/L (from 14 to 28-day LC50).

Algae: six species, 2 µg/L (Enteromorpha sp, from 5-day LC50) to 1000 µg/L; five species had some end-points with means <25 µg/L.

A marine high reliability trigger value for copper of 1.3 µg/L was derived using the statistical distribution method with 95% protection. This figure is above the converted NOEC for Mytilus edulis but below the experimental EC50 (2 µg/L) and is considered appropriate for slightly-moderately disturbed systems.


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