Commonwealth of Australia 2010
Effects on aquatic organisms
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- 50.1.1Freshwater aquatic toxicity data
- 50.1.2Marine aquatic toxicity data
- 50.1.3Chronic effects on fish and invertebrates
- 50.1.4Factors affecting the aquatic toxicity of cyanide
- 51.Water pH
- 52.Water temperature
- 53.Dissolved oxygen
- 54.Life stage, species and co-associated chemicals
- 55.Cyanide forms, complexes and other products
Effects on aquatic organismsApplicants to this assessment provided very few aquatic toxicity tests or data, and a range of literature sources were reviewed for this assessment. The effects of cyanide on fish and other aquatic organisms have been reviewed by the US Environmental Protection Agency (USEPA, 1985), the Canadian Council or Resource and Environment Ministers (CCREM, 1987), Eisler et al. (1999) and the Australia and New Zealand Environment and Conservation Council (ANZECC/ARMCANZ, 2000a). ANZECC/ARMCANZ (2000a) provide a most recent and thorough summary of aquatic toxicity data for cyanide. Data for cyanides are also listed in the very comprehensive US EPA ECOTOX database (http://www.epa.gov/ecotox/), and this source was consulted to augment the data from the above reviews, both to add more information where details were unclear and to check for more recent data that may not have been considered by them. The toxicity classification of cyanide to aquatic organisms according to the Globally Harmonized System (GHS) for Hazard Classification and Communication (UNECE, 2005) is discussed in Appendix 3. Australian water quality guidelines for the protection of aquatic ecosystems have been presented in Section 11.9.1. 50.1.1Freshwater aquatic toxicity dataAquatic toxicity data for freshwater fish, invertebrates, algae, and macrophytes have been described in Table 9.. In general, the data have been peer reviewed prior to publication (e.g. Low and Lee, 1981; USEPA, 1981, 1985; ANZECC/ARMCANZ, 2000a), and data in ANZECC/ARMCANZ (2000a) have undergone rigorous scientific review. 50.1.2Marine aquatic toxicity dataAquatic toxicity data for marine fish, invertebrates, algae and coral species have been presented in Table 9.. In general, the data presented have been peer reviewed prior to publication, and most of the data have undergone rigorous scientific review by ANZECC/ARMCANZ (2000a). Data available for anemones and corals (Chalker and Taylor, 1975; Barnes, 1985; Jones and Stevens, 1997; Cervino et al., 2003) refer to acute tests involving very short duration (1-60 minutes) exposures to cyanide in seawater, indicative of the exposure duration experienced during illegal cyanide fishing practices rather than those of standard laboratory methods. Table 9.. Summary of selected freshwater aquatic toxicity data for cyanide
Table 9.. Summary of selected marine aquatic toxicity data for cyanide
50.1.3Chronic effects on fish and invertebratesFish are a cyanide-sensitive group of aquatic organisms (Eisler, 1991; Eisler et al., 1999). Under conditions of continuous exposure for prolonged periods (early life stage, partial life cycle and life cycle studies with test durations of 28-256 d), adverse effects on fish ability at swimming and reproduction usually occur at concentrations between 5 and 7 µg CN/L and on survival between 20 and 76 µg CN/L. Other adverse effects on fish from chronic cyanide exposure include susceptibility to predation, disrupted respiration, osmoregulatory disturbances, and altered growth patterns (Table 9.). Free cyanide concentrations between 50 and 200 µg CN/L are fatal to the more sensitive fish species over time, and concentrations >200 µg CN/L are rapidly lethal to most species (USEPA, 1980). Cyanide-induced pathology in fish includes subcutaneous haemorrhaging, liver necrosis and hepatic damage. Exposure of fish to 10 µg CN/L for 9 days was sufficient to induce extensive necrosis in the liver, although gills showed no damage. Liver histopathology intensification was evident at exposure concentrations of 20 to 30 µg CN/L and exposure periods up to 18 days (Leduc, 1984). Table 9.. Symptoms of cyanide exposure in freshwater fish
Cyanide has a strong, immediate and long-lasting inhibitory effect on the swimming ability of fish (Leduc, 1984). Free cyanide concentrations as low as 10 µg CN/L can rapidly and irreversibly impair the swimming ability of salmonids in well-aerated waters (Doudoroff, 1976). This may potentially be due to respiratory and neurological effects, as observed in higher order organisms (e.g. mammals, birds). Cyanide affects fish reproduction by reducing both the number of eggs spawned, and the viability of the eggs by delaying the process of secondary yolk deposition in the ovary (Lesniak and Ruby, 1982; Ruby et al., 1986). Vitellogenin (Vtg), a glyco-lipophosphoprotein present in plasma of fish during the process of yolk formation, is synthesised in the liver under stimulation of oestrogen and subsequently sequestered in the ovary. Vitellogenin is essential for normal egg development. Exposure of naturally reproducing female rainbow trout (Oncorhynchus mykiss) to 10 µg CN/L for 12 days during the onset of the reproductive cycle produced a reduction in plasma vitellogenin levels and a reduction in ovary weight. The loss of vitellogenin in the plasma removed a major source of yolk (Ruby et al., 1986). There is uncertainty as to whether this effect of cyanide in fish is related to an effect on hormone(s) of the endocrine system. Reproductive impairment in adult bluegills (Lepomis macrochirus) occurred following exposure to 5.2 µg CN/L for 289 days (USEPA, 1980). Newly fertilised eggs are usually resistant to cyanide prior to blastula formation; however, delayed effects occurred at 60 to 100 µg CN/L, including birth defects and reduced survival of embryos and newly hatched larvae (Leduc, et al., 1982). Concentrations of 10 µg CN/L or greater have caused developmental abnormalities in embryos of Atlantic salmon (Salmo salar) after extended exposure (Leduc, 1978). These abnormalities, which were absent in controls, included yolk sac dropsy and malformation of eyes, mouth and vertebral column (Leduc, 1984). Dropsy is a distension of the abdomen, giving the fish a ‘pot belly’ appearance. This is a strong indicator of disease problems, which may include swelling of internal organs (liver, spleen or kidney), build up of body fluids (oedema, ascites), parasite problems, or other unknown cause. Dropsy is a common element in many of the serious diseases, since it is commonly associated with systemic disruption of osmoregulation due to blood-cell or kidney damage. Osmoregulatory disturbances recorded in fish exposed to 10 µg CN/L may affect migratory patterns, feeding and predator avoidance (Leduc, 1984). Increased levels of cortisol (indicating a stress response) and increased susceptibility to the pathogen Saprolegia parasitica were found in rainbow trout exposed to cyanide at 70 µg/L for 24 h (Carballo et al., 1995). Sodium cyanide has stimulatory effects on oxygen-sensitive receptors in lungfish, amphibians, reptiles, birds and mammals (Smatresk, 1986). Facultative air-breathing animals appear to rely on air breathing when external chemoreceptors are stimulated. Obligate air-breathing fish are more responsive to internal stimuli (Smatresk, 1986). Several authors (Eisler et al., 1999; Leduc, 1984; Clark, 1937) suggest a potential stimulatory effect on growth (hormesis) in animals when exposed to very low, sub-lethal concentrations; however, no test data were presented to support this hypothesis. Eisler et al. (1999) indicate that among aquatic invertebrates, adverse non-lethal effects occur with exposure to cyanide concentrations between 18 to 43 µg CN/L, and lethality between 30 and 100 µg CN/L, although some amphipod (Gammarus pulex) deaths occurred at exposure concentrations between 3 and 7 µg CN/L. USEPA (1984) noted that both freshwater and saltwater plant species (algae and macrophytes) show a wide range of sensitivities to cyanide, with the saltwater red macroalga algae Champia pulvula the most sensitive species listed (growth and reproductive effects at 11 to 25 µg/L). More recent research has shown that the saltwater diatom species Nitzschia closterium is similarly sensitive (growth rate NOEC = 10-31 µg/L). While some other algal species showed harmful effects at cyanide concentrations <100 µg/L, the available data indicate that other algal and aquatic macrophyte species are much less susceptible (effects at cyanide concentrations of 3000 to 30 000 µg/L).
However, some caution is necessary in comparing these data, as differences in the conduct of laboratory tests may have led to apparent differences between species or test conditions that are not necessarily real. For example, USEPA (1984) noted that most of the aquatic invertebrate species tested were considerably more resistant to cyanide than fishes (though Daphnia sp. and Gammarus pseudolimnaeus were comparable in sensitivity), but also observed that about half of the tests with invertebrates were under static conditions and the test concentrations were not measured, whereas many of the tests with fish were under flow-through conditions (likely to maintain more stable concentrations) and in which free cyanide concentrations were measured. 51.Water pHThe equilibrium of the chemical in water at a specific pH governs the amounts of the different forms of the chemical. These different chemical forms have different chemical properties and hence exhibit different degrees of toxicity. Cyanide equilibrium in water involves the neutral species hydrocyanic acid (HCN) and ionic form cyanide (CN). In general, low pH can increase the toxicity of cyanide (Collier and Winterbourn, 1987, Alabaster and Lloyd, 1982, CCREM, 1987). At lower pH, the proportion of HCN in solution increases (Section 23.3.1). The neutral form (HCN) is more toxic as it is able to cross biological membranes more readily than the ionic form. Solution pH levels in the range of 6.8 to 8.3 have little effect on cyanide toxicity to aquatic organisms, but toxicity is enhanced at more acidic pH (Eisler et al., 1999). 52.Water temperatureTemperature affects cyanide toxicity differently depending on the aquatic species. The toxicity of cyanide to rotifers, snails and water fleas (Daphnia spp.) increased with an increase in temperature (Cairns et al., 1978). The 48-h LC50 for the snail Nitocris sp. decreased from 13 600 µg CN/L at 5°C to 7000 µg CN/L at 25°C (toxicity increased). Similar 2-fold increases in toxicity with increasing temperature were reported for Daphnia magna, D. pulex and a rotifer. The increase in toxicity of cyanide at higher temperature was explained in part by increased metabolism of the organism (Cairns et al., 1978). In contrast, the oligochaete (Aeolosoma headleyi) showed the opposite trend with 48 h LC50 values of 9000 to 10 000 µg CN/L at 10 and 5C, compared to 120 000 µg CN/L at 15C and 160 000 µg CN/L at 20°C and 25C. Similarly, toxicity to several crustaceans was around 2 times higher at temperatures above 31C, but temperature did not appear to affect toxicity to insects or molluscs under similar conditions. Season and exercise modify the lethality of HCN to juvenile rainbow trout (McGeachy and Ludec, 1988). Cairns et al. (1978) did not report any effect of temperature on cyanide toxicity to 5 species of fish, but they did notice a variation with different species. Brown (1968) found that cyanide was more toxic to rainbow trout fry at 3C than at 13C. Smith et al (1978) examined the effects of temperature on cyanide toxicity to fathead minnows (Pimephales promelas) collected as field stock and found that juvenile fish were more sensitive at lower temperatures and at oxygen levels below 5 mg/L. The 96 h LC50 varied from 53 µg CN/L at 4°C to 143 µg CN/L at 18°C. Toxicity to several fish species was around 4 times higher (lower LC50) at 31.4C than at 26.5C (Sarkar, 1990). Ingles (1982) indicated that cyanide toxicity to fish tested increased 3-fold with a 12°C decrease in temperature. 96-hour LC50 values for rainbow trout at 6.3°C, 12.3°C and 18.0°C were 28, 42, and 68 µg CN/L, respectively. Temperature effects on algal sensitivity to cyanide are inconclusive (Cairns et al., 1978).
Substantial interspecies variability exists in sensitivity to free cyanide (Eisler et al., 1999; ANZECC/ARMCANZ, 2000a), though it is highly toxic to most species (LC50/EC50 < 1000 µg/L), as shown in Table 9. and Table 9.. Eisler (1991) concluded that fish were the most sensitive aquatic organisms tested under controlled conditions and algae and macrophytes were comparatively tolerant, but more recent data makes it clear that some species of aquatic invertebrates and algae are also very sensitive. Toxicity to freshwater and marine organisms appears broadly comparable, as evidenced by the similar marine and freshwater trigger values of 4 µg/L and 7 µg/L, respectively, derived by ANZECC/ARMCANZ (2000a) using statistical analyses of acute toxicity data. Ingles (1982) indicated that cyanide toxicity, in terms of survival duration, increases with chloride ion concentration. Eisler et al. (1999) reported that ammonia or arsenic act synergistically with cyanide.
Pablo et al. (1996) investigated the relative toxicities of cyanide (as NaCN) and iron-cyanide complexes including K3Fe(CN)6 and K4Fe(CN)6 to two Australian marine fish species (Australian bass Macquaria novemaculeata and black bream Acanthopagrus butcheri). The 96-hour LC50 values (in g CN/L) have been summarised in Table 9.. As indicated in Table 9. and Table 9., aquatic toxicity decreases with iron complexation of free cyanide. The toxicities of the iron-cyanide complexes were consistent with the significant toxic component being free cyanide, and the toxicity of the iron-cyanide complexes correlates well with free (rather than total) cyanide toxicity (Doudoroff, 1976). The difference in toxicity of ferricyanide and ferrocyanide is attributed to differences in the extent of dissociation to free cyanide, as affected by reaction thermodynamics and kinetics, photolysis and HCN losses through volatilisation (Pablo et al., 1996). Table 9.. Aquatic toxicity (LC50) to fish of metal-cyanide complexes and compounds and cyanide breakdown products
Formation of a nickel-cyanide complex markedly reduced the toxicity of both cyanide and nickel at high concentrations in alkaline pH. At lower concentrations and acidic pH conditions, nickel-cyanide solutions increase in toxicity by more than 1000 times, owing to dissociation of the metallocyanide complex to the more toxic form HCN (Towill et al., 1978). Table 9.. Comparison of toxicity data (96-h LC50 or 72-h EC50) for free and complexed cyanide (mg/L) to two Australian fish species, one marine microalga and one marine mollusc
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