Lead in freshwater and marine water

​​​Toxicant default guideline values for protecting aquatic ecosystems

October 2000

Description of chemical

Anthropogenic outputs of lead to the environment outweigh all natural sources (e.g. weathering of sulfide ores, especially galena), and lead reaches the aquatic environment through precipitation, fall-out of lead dust, street runoff and industrial and municipal wastewater discharges (USEPA 1976, Jaques 1985). Lead is generally present in very low concentrations in natural waters. In fresh waters, the main species of lead are PbCO₃ and lead-organic complexes, with very much smaller amounts of free lead ions. In marine waters, lead carbonate is the predominant form (Hart 1982).

Lead occurs in the +2 and +4 valency states, although elemental lead is relatively soluble in soft and acidic water (Fergusson 1990), and therefore, plays a significant role in the input of lead into the aquatic environment.

Summary of factors affecting lead toxicity

• Lead toxicity is hardness–dependent and a hardness algorithm is available (Table 3.4.3 of the ANZECC & ARMCANZ 2000 guidelines).
• Toxicity of lead is reduced by low solubility of many forms of lead in the natural environment, particularly in alkaline waters.
• Lead is strongly complexed by dissolved organic matter in most natural waters. Speciation measurements can account for this.
• Lead is adsorbed strongly by suspended clay, humic substances and other suspended material. Filtration and speciation measurements should account for this.
• Lead speciation in seawater is dominated by chloride complexing, which becomes negligible at salinities below approximately 6%. Hence increasing salinity reduces toxicity.
• Lead can bioaccumulate in aquatic organisms but it is generally not available at sufficient concentrations to cause significant problems.

A variety of methods are available for determining the speciation of lead in water. These include:

1. Analytical techniques, such as physical separation (e.g. (ultra)filtration, dialysis, centrifugation), voltammetry (e.g. anodic/cathodic stripping voltammetry), solvent extraction and ion exchange (Florence & Batley 1980, Botelho et al. 1994, Cheng et al. 1994, Kozelka et al. 1997)
2. Theoretical techniques, such as geochemical modelling (Florence & Batley 1980).

Bioassays are typically used to determine metal-organism interactions. These can be used in conjunction with the measured and/or predicted speciation of lead to define bioavailable lead species. The current analytical practical quantitation limit (PQL) for lead is 0.05 µg/L in fresh water and 2.00 µg/L in marine water (NSW EPA 2000).

Factors that affect the bioavailability and toxicity of lead

Solubility is the primary mechanism controlling the concentration, and hence, speciation of lead (II) in natural surface waters (Fergusson 1990). In fresh surface waters at pH < 7, the free hydrated ion (Pb²⁺) is the predominant species of dissolved lead (Stumm & Morgan 1996). Lead sulfides, sulfates, oxides, carbonates and hydroxides all have low solubility (Hem & Durum 1973). At circumneutral pH (6 to 8), lead solubility is a complex function of pH and carbonate concentration but, if pH is held constant, the solubility of lead decreases with increasing alkalinity (CCREM 1987). In more alkaline waters (pH > 8.5) containing carbon dioxide and sulfur, the solubility of lead is low (< 1 µg/L). Conversely, in acidic conditions (pH < 6) the solubility of lead increases, particularly in waters of low alkalinity (Hem 1976).

In seawater, lead speciation is dominated by chloride complexation (> 90%). The relative importance of such complexes decreases markedly with decreasing salinity, becoming negligible at salinities below approximately 6% (Fergusson 1990).

Lead (II) is strongly complexed by dissolved organic matter (DOM) in natural waters (Hodson et al. 1979, Saar & Weber 1980), and lead–DOM complexes will account for the majority of dissolved lead in natural freshwater (pH 5 to 9). Elbaz-Poulichet et al. (1984), using anodic stripping voltammetry (ASV) reported that about 20 to 30% of dissolved lead is complexed by DOM in the Gironde estuary. Capodaglio et al. (1990) reported that 50 to 70% of dissolved lead in the North Pacific was complexed by DOM. These complexes are likely to reduce lead toxicity but this has not been clearly demonstrated (Spry & Wiener 1991).

Sorption is also an important mechanism controlling the concentration of lead in natural waters (CCREM 1987). Lead is precipitated and/or adsorbed in the presence of clay suspensions (CCREM 1987), humic substances (humate) and iron, aluminium and manganese (oxy)hydroxides (Florence & Batley 1980, Dzombak & Morel 1990, Bargar et al. 1997). Lead is strongly adsorbed by humate in sediments (Corrin & Natusch 1977, Waller & Pickering 1993, Botelho et al. 1994) and these lead-humate complexes are relatively stable across a large pH range (Waller & Pickering 1993).

The uptake and toxicity of lead in freshwater organisms generally decreases with increasing water hardness and alkalinity [see reviews by USEPA (1985b), CCREM (1987), Markich & Jeffree (1994)]. For example, Davies et al. (1976) reported that the 19-month LOEC for rainbow trout (Oncorhynchus mykiss) was 4.1 to 7.6 µg/L in soft water (hardness, 28 mg/L as CaCO₃; alkalinity, 26 mg/L as CaCO₃; pH, 6.65 to 7.34). In contrast, in hard water (hardness, 350 mg/L as CaCO₃; alkalinity, 240 mg/L as CaCO₃; pH, 7.64 to 8.25) it was 18 to 32 µg/L. There is a disproportional inverse relationship between the bioaccumulation of lead and an increase in calcium concentration (Varanasi & Gmur 1978, Markich & Jeffree 1994). An exponential, inverse relationship has been shown demonstrated between water hardness and the uptake and toxicity of lead. An algorithm describing this relationship has been used to calculate a hardness-modified lead guideline value for protecting aquatic ecosystems in North America (USEPA 1995a,b).

The general belief is that the uptake and toxicity of lead is enhanced at low pH (< 6), compared to that at circumneutral pH (6 to 8) (Campbell & Stokes 1985, Wren & Stephenson 1991, Spry & Wiener 1991, Gerhardt 1994). For example, Wiener (1983) reported a ten-fold increase in the lead tissue concentration of bluegill sunfish (Lepomis macrochirus) in low pH lakes compared to that in neutral pH lakes.

Denton and Burdon-Jones (1982, 1986) found that the toxicity of lead to the banana prawn (Penaeus merguiensis), diamond-scaled mullet (Liza vaigiensis) and glass perch (Priopidichthys marianus) was reduced when salinity increased (20 to 36%).

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

Aquatic toxicology

Both acute and chronic toxicity of lead to several species of freshwater animals was greater in soft water than in hard water. At a hardness of 50 mg/L (as CaCO₃) the acute sensitivities of 10 freshwater species ranged from 143 µg/L for an amphipod to 236 µg/L for a midge (USEPA 1985b). Acute toxicities for Australian freshwater species ranged from 180 µg/L to 500 µg/L (Bacher & O’Brien 1990). Reproduction of Daphnia magna was impaired 16% by 30 µg/L lead in soft water, and 44% of trout developed spinal deformities at lead concentrations of 31 µg/L in soft water. However, in hard water, none of the rainbow trout showed deformities at concentrations of 190 µg/L (Biesinger & Christensen 1972). Freshwater algae were affected by concentrations of lead above 500 µg/L, based on data for four species (USEPA 1986). Bioconcentration factors for four species of invertebrates and two species of fish ranged from 499 to 1700 (USEPA 1985b).

The acute toxicity for 13 marine animal species ranged from 315 µg/L (mummichog) to 27,000 µg/L (soft-shell clam). Fewer data are available for chronic toxicity in marine waters. Unacceptable effects for mysids were observed at 37 µg/L, and macroalgae were affected at 20 µg/L (USEPA 1986).

Freshwater guideline

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 just 19 data points covering five taxonomic groups. Data were corrected to low hardness (30 mg/L CaCO₃) and amended to no observed effect concentration (NOEC) equivalents using an adaptation of the method of van de Plassche et al. (1993), and are summarised below as geometric means of NOECs.

Fish: four species, 5.65 µg/L (Lepidomeda vittata) from maximum acceptable toxicant concentration (MATC) reproduction to 43 µg/L (Salmo salar, from chronic LC50).

Amphibian: one species, Ambystoma opacum, 68 µg/L (from LC50).

Crustaceans: two species, 5.1 µg/L (Gammarus pseudolimnaeus, from LC50 and LOEC) to 19.5 µg/L (D. magna, from EC50 and NOEC reproduction).

Insects: one species, Tanytarsus dissimilis, 28 µg/L, from LC50.

Molluscs: oine species, Dreissena polymorpha, 28 µg/L, from LC50.

This figure was equal to the lowest single NOEC value but was less than the geometric mean for this species, and is considered acceptable for slightly to moderately disturbed ecosystems.

Marine guideline

The screened marine data for lead comprised 25 data points covering four taxonomic groups, as follows.

Crustaceans: one species, Mysidiopsis bahia, 29 to 51-day NOEC, reproduction, 25 µg/L.

Molluscs: one species, Perna viridis, 7-day LC50, 4400 to 4520 µg/L (giving a NOEC of 880 to 904 µg/L).

Annelids: two species, 28-day LC50, 840 to 7550 µg/L; 183 to 274 day LOEC, reproduction, 20 µg/L, converting to NOEC of 8 µg/L.

Algae: three species, 10-day EC50 (one species), 3110 to 7940 µg/L; 14-day MATC (two species), reproduction, 16 to 54 µg/L, converting to NOEC of 8 to 27 µg/L.

References

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Biesinger KE & Christensen GM 1972. Effects of various metals on survival, growth, reproduction and metabolism of Daphnia magna. Journal of the Fisheries Research Board of Canada 29, 1690-1700.

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