Zinc in freshwater

​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

Zinc can enter the environment from both natural processes (e.g. weathering and erosion) and anthropogenic (e.g. zinc production, waste incineration, urban runoff) processes (CCREM 1987). Zinc is an essential trace element required by most organisms for their growth and development. It is found in most natural waters at low concentrations (Table 8.3.2 of the ANZECC & ARMCANZ 2000 guidelines).

Summary of factors affecting zinc toxicity

  • Zinc is an essential trace element required by many aquatic organisms.
  • Zinc toxicity is hardness–dependent (also alkalinity) and a hardness algorithm is available (Table 3.4.3 of the ANZECC & ARMCANZ 2000 guidelines). Toxicity decreases with increasing hardness and alkalinity (Holcombe & Andrew 1978, Mount 1986).
  • Levels of dissolved organic matter found in most freshwaters are generally sufficient to remove zinc toxicity but often not in very soft waters. Speciation measurements can account for this.
  • Zinc forms complexes with dissolved organic matter, the stability of which depends on pH. Organic complexation is common in marine waters.
  • Zinc is adsorbed by suspended material. Filtration and speciation measurements should account for this. There is conflicting evidence on its bioavailability after adsorption.
  • Zinc toxicity generally decreases with decreasing pH, at least below pH 8. Trends are complex above pH 8.
  • Zinc uptake and toxicity generally decreases as salinity increases.


In natural waters at pH≤8.5, the predominant species is the +2 valency state (Stumm & Morgan 1996). In estuarine waters, at neutral pH, the predominant species of zinc is Zn2+, whereas at higher pH (pH≥8), in the open sea, the hydrolysed species, ZnOH+ and Zn(OH)2, become the major species (Young et al. 1980, Bervoets et al. 1996).

A variety of techniques are available for determining the speciation of zinc in water. These include:

  1. Analytical techniques, such as physical separation (e.g. (ultra)filtration, dialysis, centrifugation), polarography, voltammetry (e.g. anodic/cathodic stripping voltammetry), ligand competition and ion exchange (Cheng et al. 1994, Apte & Batley 1995, Vega et al. 1995)
  2. Theoretical techniques, such as geochemical modelling (Wilson 1978, Bervoets et al. 1996, Stumm & Morgan 1996).

Bioassays are typically used to ascertain metal-organism interactions. These can be coupled with measured and/or predicted speciation calculations to determine the bioavailability of various zinc species. The current analytical practical quantitation limit (PQL) for zinc is 0.2 µg/L in fresh water (NSW EPA 2000).

Factors that affect the toxicity of zinc

It is generally considered that Zn2+ is the form of zinc primarily responsible for eliciting a toxic response in aquatic organisms. Typically, inorganic and organic complexes ameliorate the uptake and toxicity of zinc by reducing the concentration of Zn2+.

A number of studies have established the uptake and toxicity of zinc in aquatic organisms decreases with increasing water hardness (e.g. Mount 1966, Holcombe & Andrew 1978, Bradley & Sprague 1985, Everall et al. 1989). Holcombe and Andrew (1978) determined a zinc toxicity (LC50) in soft water (hardness, 44 mg/L as CaCO3) of 0.76 and 2.4 mg/L to rainbow trout (Oncorhynchus mykiss) and brook trout (Salvelinus fontinalis), respectively. In hard water (hardness, 170 mg/L as CaCO3), the corresponding toxicity values were 1.9 mg/L for the rainbow trout and 5.0 mg/L for the brook trout. The difference between the alkalinity (43 mg/L as CaCO3) and pH (7.35) of the two test waters was negligible. The study of Holcombe and Andrew (1978) also indicated that an increase in alkalinity and pH further ameliorated zinc toxicity to the two trout species.

An exponential, inverse relationship has been shown to exist between water hardness and the uptake and toxicity of zinc. An algorithm describing this relationship has been used to calculate a hardness-modified zinc guideline value for protecting aquatic ecosystems in North America (USEPA 1995a,b).

There is a consensus of opinion that below pH 8 zinc toxicity decreases with decreasing pH (Holcombe & Andrew 1978, Bradley & Sprague 1985, Harrison et al. 1986, Everall et al. 1989, Roy & Campbell 1995). At low pH (i.e. pH 4) an increase in toxicity may be observed due to increased acidity (Fromm 1980). Conflicting results have been reported for zinc toxicity at higher pH (8 to 9) (Farmer et al. 1979, Bradley & Sprague 1985, Everall et al. 1989).

Redox will have little direct influence on zinc speciation, however, in reducing waters, and in the presence of sulfur, insoluble ZnS(s) will reduce the dissolved zinc concentration (Young et al. 1980).

Zinc forms complexes with natural DOM, the stability of which are dependent on the pH, the aqueous concentration of zinc and the presence and concentration of other ions in the waters (Florence & Batley 1977). Alkaline conditions favour the formation of Zn-DOM, ZnOH+ and ZnCO3; the latter complex being more prevalent in waters of increased alkalinity (Wilson 1978). In estuarine waters, recent studies suggest that zinc-DOM complexes comprise upwards of 50% of total dissolved zinc (van den Berg et al. 1986, 1987, Muller & Kester 1991). Bruland (1989) has shown that > 98% of dissolved zinc in the surface waters of the North Pacific is complexed by natural organic ligands.

There have been few studies that have investigated the uptake and/or toxicity of zinc in the presence of DOM. Vercauteren and Blust (1996) found that the bioavailability of zinc to the common marine mussel Mytilus edulis, was reduced in the presence of five organic ligands.

Anderson and Morel (1978) demonstrated that organic complexation of natural background levels of zinc in coastal lagoons can limit the growth of diatoms. Morel et al. (1994) postulated that natural oceanic zinc levels might have an effect on global primary production and the carbon cycle.

The removal of zinc from solution via adsorption processes is an important process in natural waters (CCREM 1991). Zinc can sorb to iron, aluminium and manganese (oxy)hydroxides (Lee 1975, Dzombak & Morel 1990), clay minerals (USEPA 1979d) and colloidal organic matter (Tessier et al. 1996). In acidic waters (pH <6) little zinc is expected to (CCREM 1991). As salinity increases, adsorptive capacity is expected to decrease (James & McNaughton 1977).
The uptake and toxicity of zinc decreases with increasing salinity (Nugegoda & Rainbow 1989, Hamilton & Buhl 1990, Bervoets et al. 1996).

Jarvinen and Ankley (1999) report data on tissue residues and effects for zinc for eight freshwater species and six marine species. It is not possible to summarise the data here but readers are referred to that publication for more information.

Aquatic toxicology

USEPA (1987b) compiled acute toxicity values of zinc for 43 freshwater species. At a hardness of 50 mg/L, the concentrations ranged from 51 µg/L to 81,000 µg/L. Bacher and O’Brien (1990) found the acute toxicities for Australian freshwater species ranged from 140 µg/L to 6900 µg/L, and Skidmore and Firth (1983) found a range of 340 to 9600 µg/L for ten Australian species. Zinc was found to bioaccumulate in freshwater animal tissues 50 to 1130 times but bioaccumualtion is not generally considered a problem for zinc.

Freshwater guidelines

For freshwater guideline derivation, only the chronic data that were linked to pH and hardness measurements were considered and further screened for quality and other factors. This reduced the dataset to around 85 data points. These were adjusted for uniform lower hardness (30 mg/L as CaCO3) and other end-points adjusted to NOECs using the method adapted from van de Plassche et al. (1993). The NOEC values from six taxonomic groups were as follows (pH range 6.75 to 8.39):

Fish: 11 species, 24 µg/L (Oncorhynchus tshawytscha; from LC50) to 1316 µg/L (Ptylocheilus oregonensis; from LC50); seven species had geometric means < 250 µg/L and a measured NOEC of 38 µg/L was reported for Pimephales promelas.

Amphibians: one species, Ambystoma opacum, 180 µg/L (from LOEC).

Crustaceans: three species, 5.5 µg/L (C. dubia; from LC50) to 25.3 µg/L (C. dubia), plus a figure of 18,480 for the crayfish Orconectes virillis).

Insect: one species, Tanytarsus dissimilis, 5 µg/L (NOEC).

Molluscs: three species, 54 µg/L (Dreissena polymorpha) to 11,200 µg/L (Velesunio ambigua), a NOEC of 487 µg/L was measured for Physa gyrina.
Annelid: one species, Limnodrilus hoffmeisteri, 560 µg/L (from LC50).

The geometric means for zinc were distinctly bimodal with two values at least 9.4 times the next highest. However, all the data fitted the model and they were not excluded. The trigger value is above the lowest measured NOEC for an insect and the recalculated NOEC for C. dubia (from chronic LC50 of 27.5 mg/L).

However, given the essential nature of zinc and the fact that the chronic end-points are NOECs, the risk is low and the 95% protection level is considered acceptable for slightly to moderately disturbed systems.

A freshwater high reliability trigger value of 8 µg/L was calculated for zinc using the statistical distribution method with 95% protection. This applies at hardness of 30 mg/L of CaCO3.


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