Ecotoxocicological assessment of antifouling biocides and non-biocidal antifouling paints

2. Copper

2.1 Copper concentrations measured in the vicinity of pleasure craft harbours
2.2 Transformation and bioavailability of copper in water and sediment
2.3 Release and sequestration of copper in sediments
2.4 Bioaccumulation and aquatic toxicity
2.4.1 Bioaccumulation
2.4.2 Toxicity to aquatic organisms
2.5 Assessment of copper

2.1 Copper concentrations measured in the vicinity of pleasure craft harbours

Denmark
In the Egå Marina at Århus Bay, the copper content in the harbour sediment was 150-600 mg/kg dry weight at a distance of 5-10 m from discharge from consolidated areas, decreasing to 53-120 mg/kg dry weight at 30 m from discharge (Jensen and Heslop 1997a). By way of comparison, the copper content in sediment in Århus Bay was 25-50 mg/kg dry matter. Finally, in the same investigation, water concentrations of copper of 2.4 µg/L were measured in Studstrup and of 13 µg/L in the pleasure craft harbour of Marselisborg but these analyses are stated to be somewhat doubtful.

The copper content in harbour sediments from other localities in the area has also been analysed. The highest concentrations were found at the slipways in Bønnerup harbour (7,000-8,000 mg/kg dry weight), which is a combination of a pleasure craft and fishing harbour, and in Århus fishing port (1,600-2,400 mg/kg dry weight). The copper concentrations in the basins were 15-70 mg/kg dry weight in Bønnerup harbour and 100-400 mg/kg dry weight in Århus fishing port. In the sediment from Ebeltoft, Grenå and Hov Bedding, the concentrations were 280, 490 and 1,200 mg/kg dry weight, respectively (Jensen and Heslop 1997b).

The county of Funen has measured copper contents in sediment from 5 to 110 mg/kg dry weight in harbours (The County of Funen 1999). From the Little Belt, dated sediment cores have been analysed so that the temporal development of the copper content might be assessed. The measurements in the sediment cores showed a significantly increasing content of copper in the vicinity of Als, an upward trend at four stations and constant/varying concentrations at four other stations. The copper content in the cores varied from 19 to 46 mg/kg dry weight.

Sweden
In 1990 and 1993, the copper concentrations in water, sediment and aquatic plants were measured in the skerries of Stockholm (Greger and Kautsky 1990 and 1993, cf. Bard 1997). The measurements showed a significantly higher content of copper in the sediments in the vicinity of pleasure craft harbours and areas with heavy pleasure craft traffic. Copper concentrations of up to 1,3000 mg/kg dry weight were measured in sediments. Compared to less contaminated areas, increased copper concentrations were also found in aquatic plants. Similar measurements were performed in the vicinity of the Bullandö Marina, which is also situated in the skerries of Stockholm (Öhrn 1995, cf. Bard 1997). In April, 1993, before the start of the sailing season, the copper content in the water was 0.8-1.0 m g/L while, in June, it was 3.0-3.8 m g/L. The copper content in the sediment at the Bullandö Marina was only slightly increased when compared to the reference stations, at which the copper content was 30 mg/kg dry weight.

France
Measurements performed by the French authorities in the Arcachon Bay at the Atlantic coast from 1979-1991 showed an increase in the copper content in oysters (Claisse and Alzieu 1993). The increase was significant at two of four stations from 1982 to 1991. The increase in the copper content in oysters coincides with an increased consumption of copper-based antifouling products when the use of organotin TBT was regulated in 1982. The increase in the copper content was highest and significant in the oysters from the two stations in the inner bay. The increase was less and not significant at the stations in the outer bay, which may be explained by the fact that the rate of water renewal at these stations is higher than in the inner bay. The French measurements are unique because of the long time series and the extensive measuring programme carried out. Several conditions may influence the accumulation of copper in organisms and the direct relation between an increased copper content in water and an increased content of copper in oysters is not stated in the French studies.

Background concentrations
Compared to the above stated concentrations, the background concentration of copper is given as 25-35 mg/kg dry weight in Danish sediments and as 0.5-1.5 m g/L in seawater (Madsen et al. 1998). Swedish studies give a copper content of 0.3-0.8 m g/L in water from the Baltic Sea and of 0.2 m g/L in water from the Kattegat. The Swedish background values for sediment are given as 10-40 mg/kg dry weight in the Baltic Sea (Debourg et al. 1993).

In harbours and neighbouring waters, concentrations above normal of copper have thus been found in both sediments and water samples, i.e. in pleasure craft harbours up to a factor of 30 times the background concentration in sediments and up to a factor of 10-15 times the background concentration in water.

2.2 Transformation and bioavailability of copper in water and sediment

Bioavailability
Contrary to organic compounds used in antifouling products, metals are not degradable. In water, copper will occur dissolved in the water as well as sequestrated to particles. Copper may, however, occur in different forms (species) depending on e.g. the salinity, pH, content of organic matter, etc. of the water. The speciation of the copper decides whether live organisms can take it up (whether it is bioavailable) and thereby whether copper is toxic to the organisms.

It is often accepted that primarily the free copper ions (Cu²⁺) may pass cell membranes and thus constitute the bioavailable and toxic part of copper (Campbell 1995). It has, however, been demonstrated that other copper ions and lipid-bound copper may also pass cell membranes and may thus also be bioavailable (Allen 1993).

Sequestration
In seawater and fresh water, it is a well-known fact that sequestration of copper to organic substances is predominant (Bruland et al. 1991), which typically reduces the bioavailability of copper (Lewis 1995). It is, however, not that simple as there are differences in the sequestering properties of organic substances in relation to copper. E.g., Garvey et al. (1991) have demonstrated that humic acid reduces the toxicity of copper while fulvic acid does not have a similar effect. In all probability, the sequestration of copper to organic substances is very specific (Wells et al. 1998). It has been demonstrated that planktonic algae can excrete organic substances binding copper (Brand et al. 1986, cf. Wells et al. 1998). Planktonic algae that are exposed to increased copper concentrations may excrete such copper-binding substances (ligands) thereby reducing the bioavailability and potential toxic effect of copper (Wangersky 1986, cf. Paulson et al. 1994). The formation of colloids and subsequent aggregation, which eliminates copper from the water phase and transports it to the sediment, is considered another effect of the organic ligands (Wells et al. 1998).

Sedimentation, speciation and bioavailability
The transportation of copper to sediments will typically proceed via sedimentation of copper built into or adsorbed to particles (micro algae, clay particles, etc.). In open waters, the sedimentation of copper will primarily be controlled by the sedimentation of planktonic algae (Wangersky 1986, cf. Paulson et al. 1994), to which copper is sorbed and/or built in. In the sediment, a large number of chemical and biological transformations of importance to the speciation of copper will take place, including oxidation/reduction, dissolution/leaching and sequestration. The transformations will be controlled by sediment type (i.a. grain size and content of organic matter), digging and filtering activity of sediment-living invertebrates (bioturbation) and the oxygen conditions in the water and in the sediment.

Speciation of copper in sediments is controlled by dynamic and reversible processes (Calmano et al. 1990). E.g., copper sequestrated to reduced compounds (organic matter and sulfides) may be released from the sediment to the above water due to oxidation as a result of resuspension or bioturbation (Petersen et al. 1997; Ciceri et al. 1992; Westerlund et al. 1986), or a redistribution may take place sequestrating copper in oxidized compounds instead (e.g., ferric or manganese oxides and hydroxides). These compounds are considered unstable while sulfides and organic substances are characterized as more stable (Förstner et al. 1990; Calmano et al. 1990).

In anoxic sediments, e.g., in fine-grained sediments with high content of organic matter, copper will typically sorb to sulfides and organic matter while, at good oxygen conditions, copper will typically be sequestrated to compounds like ferric oxides, manganese oxides and hydroxides. Metal sulfides are recalcitrant but relatively easily and rapidly oxidized at good oxygen conditions (Förstner 1985).

The bioavailability of copper in sediments is an extremely complex phenomenon that does not depend only on the speciation and the sediment but also on the physiology and food choice of the exposed organisms (Slotton and Reuter 1995). It has been demonstrated that the bioavailability may be specific for individual species and that variations occur within the same species related to age, sex and size of the organism (Lewis 1995). Furthermore, it has been shown that the organisms take up more easily metals sorbed to easily digested food than metals sorbed to food hard to digest (Wang and Fisher 1996). Digestive enzymes in the intestine ensure a high utilization of the food (Forbes et al. 1998), which may also result in an increased uptake of copper from sediment.

Assessment of bioavailability
Increased concentrations of metals in aquatic sediments are widespread and the authorities must often consider whether the increased concentrations imply a risk of adverse effects on the ecosystem. Unfortunately, this problem is versatile as the bioavailability of metals varies a great deal in different sediments (Luoma 1989).

In attempts to predict the bioavailability of metals in sediments on the basis of chemical analyses, various extraction and fractionation guidelines have been developed for analyses of copper sequestrated to carbonates, manganese oxides, ferric oxides and organic substances (e.g., Förstner 1985). The problem in these extraction and fractionation guidelines is, however, to interpret which species are bioavailable. On the basis of investigations showing a correlation between the cadmium concentration in pore water in sediment and the acute toxicity of cadmium-added sediment to an amphipod (crustacean living in holes in the sediment), the assumption that the content in the pore water represented the bioavailable part of cadmium was proposed (Ankley et al. 1994).

In similar investigations of the effects of cadmium on other amphipods, Di Toro et al. (1990 cf. Ankley 1996) have demonstrated that the acute toxicity of cadmium may be predicted on the basis of the content of acid volatile sulfide (AVS). AVS is the fraction of sulfide in the sediment that is extractable with cold hydrochloric acid and is a measurement for the capacity of the sediment to sequestrate metals. If the sequestering capacity is exceeded, the concentration of cadmium in the sediment is increased and the amphipods die. Attempts have been made to use AVS for determining the bioavailability to amphipods of copper in sediments (Ankley et al. 1993, cf. Ankley 1996). AVS significantly overestimated the bioavailability of copper, which was explained by the presence of another sequestration phase than AVS.

The concept is based on the assumption that only the content in the pore water is available combined with a steady state consideration. This assumption cannot be expected to apply to sediment reworkers that swallow whole sediment particles and have digestive enzymes in the intestine for degradation of organic substances. Furthermore, the AVS method is limited in as much as it was developed to determine only the actual bioavailable fraction of metals and thus does not give a measurement for the potentially bioavailable fraction that may eventually become bioavailable, e.g., in relation to a change in oxygen conditions.

No simple method based on chemical analysis has thus yet been found with which you can assess how large a part of the copper - especially in sediments - that is bioavailable and it is questionable under which conditions (sediment type, oxygen conditions) and for which organisms, the AVS method is valid.

2.3 Release and sequestration of copper in sediments

Dredging and dumping
In connection with resuspension of sediments, it was demonstrated that a considerable part of the sorbed copper may be released from the sediment. In laboratory experiments under natural conditions, it was found that up to 2% of the particle-sequestrated copper may be released to the water at resuspension (Petersen et al. 1997). An investigation of sediments at dumping sites at Cleveland Bay before and after a dredging and dumping concludes that copper in the sediment is sequestrated in labile fractions, which are potentially bioavailable and which are easily spread at resuspension (Reichelt and Jones 1994).

Measured release
It has been demonstrated that metals (especially copper) may be released from sediments to the water above sediments with oxidized surface (Luoma 1989). Release of copper from sediments was observed at the North American coast (Boyle et al. 1981 cf. Luoma 1989), in the North Sea (Kremling 1983, cf. Luoma 1989) and in several places in coastal areas (Windom et al. 1983, cf. Luoma 1989). Fencing experiments have shown that the copper release is larger from copper-contaminated sediments than from uncontaminated sediments (Hunt and Smith 1983).

Bioturbation
Bioturbation in sediments may be of great importance to remobilization of metals in the sediment. Sediment-living animals are characterized in relation to their search for food. Sediment reworkers swallow sand, mud and water without previous separation. The organic content in sediment is low compared to other types of food. In order to compensate for this, the sediment reworkers have to consume large amounts of sediment, some ingest 8-10 times their body weight a day. Sediment reworkers typically rummage about a lot in the sediment, which may mobilize buried metals. Suspension feeders feed on particles, which they filter from a current that they create between the water above the sediment and the sediment itself thereby increasing the exchange of substances above the sediment-water surface. All in all, the animals increase the contact between the sediment and the above water. Considering that the net deposition in marine sediments is only a few millimetres a year, the animals may contribute to bringing up old sediment to the surface and new sediment down to underlying layers. Peterson et al. (1996) found that bioturbation could significantly increase the bioavailability of metals in sediments through oxidation of sulfide compounds. They found that metal/sulfide complexes were relatively unstable towards the oxidation taking place in connection with bioturbation.

The replacement of sulfide-containing water by oxygen-containing water will also remobilize sulfide-sequestrated metals (Emerson et al. 1984, cf. Förstner et al. 1990) as the oxygen content of the water above the sediment is of great importance to the sequestration and release of metals from sediments. Measurements showed that, during summer periods with poor oxygen conditions in the harbour at Corpou Christi Bay, cadmium was sequestrated to sulfides while measurements showed a release during winter months with good oxygen conditions (Holms et al. 1974, cf. Förstner et al. 1990). There is thus no immediate reason to suppose that copper sequestrated in sulfides may not become bioavailable on a long view.

2.4 Bioaccumulation and aquatic toxicity

Copper is a micro-nutrient that live organisms need in small doses. Higher animals like fish can regulate the content of copper in their organism and, to some extent, they can accumulate copper in the lever but not in the muscles. If copper exists in the surroundings or in the food in very low concentrations, an accumulation may be the result of the organism utilizing copper as a nutrient. The interpretation of bioconcentration factors (BCF values) for an essential micro-nutrient like copper is thus difficult and no information is available in the investigations quoted on concentrations of copper and the requirements for copper of the organisms used. In short-term studies with algae (½-2 days), BCF values were measured at 1-40. In long-term studies with insects and mussels, the BCF values were considerably higher: In a 28-day study with mosquito larvae - in all probability in sediment - a BCF value of 5,830 was found; furthermore, BCF values of 5,000-10,000 were found in mussels during a period of 2-3 years (AQUIRE 1999). BCF values between 400 and 90,000 have been found in plankton and some lower organisms (Debourg et al. 1993).

2.4.2 Toxicity to aquatic organisms

Aquatic organisms
Table 2.1 gives an overview of the toxicity of copper to various groups of aquatic organisms measured in single-species laboratory tests. Table 2.1 illustrates that copper is very toxic with effect concentrations from only a few micrograms of copper per litre.

In Denmark, quality criteria have been specified for copper in fresh water and seawater of 12 m g/L and 2.9 m g/L, respectively (The Danish Ministry of Environment and Energy, 1996). It is, however, stated that the criteria are based on data that have not finally been quality assessed. On the basis of 65 single-species laboratory tests with marine organisms, a PNEC value for copper has been calculated at 5.6 µg/L (Hall and Anderson 1998). The calculation method used is based on the distribution of the sensitivity of the organisms tested, and the calculated PNEC value theoretically protects 95% of the species with 95% confidence. This is, however, twice the lowest NOEC value in Table 2.1 (0.0031 mg/L = 3.1 µg/L).

Ecosystem studies
Effects on natural planktonic algae have been measured at only a few micrograms of copper per litre. Chronic effects of copper on planktonic algae in marine ecosystem modelling were demonstrated from 1 µg/L (Gustavson et al. 1999). Comprehensive and well-documented experiments with micro algae and copper (Brand et al. 1986) show that, even at very low concentrations, copper may inhibit the reproduction of algae. In these studies, the effect of copper on 38 different clones of marine planktonic algae was examined in water, in which the metal-chelating properties were known. In this scientific article, the toxic effect of copper on the reproduction is related to the activity of free copper ions and it is concluded that copper may inhibit the reproduction of sensitive algal species even in uncontaminated waters where the copper concentration is low (0.1-0.2 µg/L). These studies distinguish themselves i.a. by relating the effect of copper to the activity of the free copper ions in the water and not only to the total copper concentration as so many other studies do.

Swedish investigations have shown copper concentrations of up to 3 µg/L in the vicinity of pleasure craft harbours in areas in which the background concentration of copper was 0.8-0.5 µg/L. At the actual copper concentrations, no effects on planktonic algae were found (Wängberg et al. 1995).

Bottom-living organisms
The results from the tests with organisms living in the sediment and at the bottom are presented in Table 2.2.

The three studies, in which the concentration is given in mg/L, may have been conducted in water without sediment. The other studies indicate that copper in sediment may cause effects on sediment-living animals at concentrations exceeding 100 mg/kg (Table 2.2). This is well over twice as much as the highest of the background concentrations stated but much lower than the concentrations measured in harbour sediments.

2.5 Assessment of copper

Copper is an element and is thus not degradable. Copper can be "removed" from the aquatic environment by sorbing to and being buried in sediments outside the reach of organisms. Seen in a geological time perspective, large amounts of heavy metals have been discharged into the sea without causing serious ecotoxic effects as the sequestration of metals to the sediment has prevented this.

In the aquatic environment, copper will sorb to inorganic and organic substances and particles. These sequestering conditions contribute to the occurrence of various species of copper. It is uncertain which species are bioavailable, and no reliable measuring methods for assessment of the size of the bioavailable fraction are available. Furthermore, the bioavailability of copper is not constant and must be view in different time perspectives. A differentiation must thus be made between the actual and the potential bioavailability. The actual bioavailability will typically be considerably less than the potential bioavailability. Furthermore, bioavailability is species specific and may also depend on physiology, nutrition, age, size and sex of the organisms in question.

A permanent immobilization of copper can only occur at sequestration to particles and subsequent sedimentation on sediments with poor oxygen conditions with a permanent presence of sulfides. In reality, such conditions only exist in areas without resuspension, i.e., without bioturbation (macro fauna) and fishery with bottom trawl. The extension of these sediment types in Denmark is limited to a few holes in i.a. the archipelago south of Funen. Copper sorbed to particles that settle on sediments rich in oxygen with bioturbation will probably stay in the biological systems for many year. In deep waters, nutrients and trace metals, including copper, stay in the water phase as the particles attain to transformation in the water column before they reach the surface of the sediment.

Harbour sediments are typically anoxic and have a high content of sulfides which will bind copper. Therefore, copper is expected to be relatively strongly sequestrated in harbour sediments. A release from the sediment at resuspension induced by e.g. the propellers of ships can, however, not be excluded. At regular intervals, the sediments in the harbours are dredged and the material is dumped at selected localities. Copper may be released at dumping and, typically for dumping sites in Denmark, the sediment will subsequently be spread by current and wave action. Stable dumping sites are difficult to find in Denmark and copper in the harbour sediments must be expected to be spread over large areas in connection with dumping.

The toxicity of copper is dependent on the speciation and the bioavailability of copper in the water. The fact that copper is a micro-nutrient combined with the fact that the content of metal chelating substances may greatly vary in time and space and that the sensitivity of different species varies much, make it very difficult to compare different investigations. The concentrations, in which effects are measured in laboratory tests, are generally higher than the background concentrations stated for copper in the environment but concentrations measured in and in the vicinity of harbours are at the same level as or higher than concentrations in which effects have been measured. The organisms that are most sensitive to copper are algae and crustaceans and, in ecosystem tests of the sensitivity of algae, effects were measured at copper concentrations on the same level as background concentrations.