Detergents 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

Detergents (or surfactants) are complex mixtures containing a variety of ingredients, particularly surface-active agents (surfactants), builders, bleaches and additives, blended for specific performance characteristics (Hennes-Morgan & de Oude 1994).

Surfactants are usually categorised into three groups, anionic, non-ionic and cationic. Anionic surfactants comprise such common groups as linear alkylbenzene sulfonates (LAS) and alkyl ethoxylated sulfates (AES). Non-ionic surfactants include alcohol ethoxylates (AE) and alkylphenol ethoxylates (APE). Cationic surfactants comprise quaternary ammonium compounds.

The most common source of surfactants in aquatic environments is from sewage treatment plants but with varying rates of degradation, the composition of surfactants in the effluent can be very different from that in the incoming wastewaters (Hennes-Morgan & de Oude 1994).

Measurement of surfactants

The standard measurement of surfactants in water is with the cationic dye methylene blue, which analyses the sum of methylene blue active substances (MBAS). MBAS measures anionic sulfonates but not AES, and is a rapid screening aid. Chromatography — gas chromatography (GC) or high performance liquid chromatography (HPLC) — can allow for more specific analysis. Analysis of bismuth active substances (BiAS) is used as a screen for non-ionic surfactants (Hennes-Morgan & de Oude 1994).

Fate in the environment

Anionic surfactants are rapidly biodegraded in sewage treatment processes, and usually > 90% is lost (Hennes-Morgan & de Oude 1994). LAS is also removed by mineralisation and biodegradation, which occurs more rapidly at higher temperature. Biodegradability generally increases with increasing chain length, although this is not so for AES surfactants (Hennes-Morgan & de Oude 1994).

For non-ionic surfactants biodegradation is not affected by chain length and is reduced above 20 ethoxylate units. Linear chains are degraded more rapidly than branched chains (Hennes-Morgan & de Oude 1994). Again, AE is substantially removed during sewage treatment.

Quaternary ammonium compounds generally adsorb strongly to suspended material and form complexes with anionic compounds. Effective degradation in sewage treatment requires prior acclimation (Hennes-Morgan & de Oude 1994).

Alkylphenol ethoxylates have been implicated in causing estrogenic effects in aquatic organisms (Nimrod & Benson 1996). It should be noted, however, that this is a rapidly developing field and a wide range of other chemicals may also be exerting such effects (refer to Section 8.3.7.21 of the ANZECC & ARMCANZ 2000 guidelines on endocrine-disrupting chemicals).

Normalisation of toxicity data

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Toxicities of detergents can vary widely with species and chemical (Lewis 1990). In order to allow detergent data to be interpreted for different detergents within the same group, the toxicity data can be normalised for a specific alkyl chain length or a specific number of ethoxylate (EO) groups, according to the method of Feijtel & van de Plassche (1995). Normalisation was carried out for short-term toxicity in the absence of equations for chronic toxicity.

For AE, EC50 is calculated using the equation:

log (1/EC50) = 0.87 log Kow + 1.13 (Konemann 1981).

For LAS and AES, EC50 is calculated using the equation:

log (1/EC50) = 0.63 log Kow + 2.52 (Saarikoski & Veluksela 1982).

Kow is calculated for the normalised structure with specified increment for each EO or alkyl group (Feijtel & van de Plassche 1995).

Aquatic toxicology

The aquatic toxicity of surfactants varies widely but normalisation, as described above (Feijtel & van de Plassche 1995), assists in interpreting and using the available data.

The normalised Dutch data were used for guideline calculation.

There have been a number of reviews and risk assessments of surfactants (Kimerle 1989, Lewis 1990, 1991, Dorn et al. 1993).

Linear alkylbenzene sulfonates (LAS)

No observed effect concentrations (NOECs) listed below are geometric means normalised to an alkyl chain length of C11.6.

Freshwater fish: five species, 250 to 3200 µg/L.

Freshwater crustaceans: two species, 1400 to 3200 µg/L.

Freshwater insects: two species, 2800 to 3400 µg/L.

Freshwater mesocosms: only one study, by Guhl and Gode (1989), meets the Organisation for Economic Co-operation and Development (OECD) requirements. This gave a NOEC of 300 µg/L, which confirms the guideline value derived from laboratory studies.

Freshwater algae: six species, 80 to 15,000 µg/L.

Marine fish: one species, Limanda yokahamae, 50 µg/L.

Marine crustaceans: one species, Mysidiopsis bahia, 120 µg/L.

Marine mussels: one species, 25 µg/L.

Alkyl ethoxylated sulfates (AES)

NOECs listed below are normalised for an alkyl chain length of C12.5 and number of EO groups of 3.4, but there was little change from original figures.

Freshwater fish: one species, Pimephales promelas, 870 to 1600 µg/L.

Freshwater crustacean: one species, Daphnia magna, 1100 to 1500 µg/L.

Freshwater rotifer: one species, Brachionus calyciflorus, 360 to 1400 µg/L. The geometric mean is 795 µg/L.

Freshwater algae: two species, 730 to 5000 µg/L.

Freshwater mesocosms: No published data were reported.

Alcohol ethoxylates (AE)

NOECs listed below are normalised to an alkyl chain length of C13.3 and EO of 8.2.

Freshwater fish: two species, 720 to 1500 µg/L.

Freshwater crustaceans: two species, 590 to 860 µg/L.

Freshwater rotifers: one species, Brachionus calyciflorus, 1300 µg/L.

Freshwater algae, diatoms and blue–green algae: six species, 200 to 8700 µg/L.

Freshwater mesocosms: 4 NOEC data for multiple species tests were 80, 80, 320 and 330 µg/L, although replication was insufficient to meet OECD (1992a) requirements. Normalised data were 380, 380, 320 and 1570 µg/L.

Marine fish: one species, Fundulus heteroclitus, 4800 µg/L.

Marine crustaceans: two species, 2700 to 48,000 µg/L.

Marine molluscs: one species, Mytilus edulis, 5500 µg/L.

Guidelines

Linear alkylbenzene sulfonates (LAS)

A high reliability freshwater trigger value of 280 µg/L was derived for LAS (normalised data) using the statistical distribution method with 95% protection. If the freshwater figure were to be adopted for marine systems, the trigger value would be greater than data derived from all four marine species. Hence, a marine low reliability trigger value of 0.1 µg/L was derived for LAS (normalised data) using an assessment factor (AF) of 200. This should only be used as an indicative interim working level.

Alkyl ethoxylated sulfates (AES)

A high reliability freshwater trigger value of 650 µg/L was derived for AES (normalised data) using the statistical distribution method with 95% protection. In the absence of marine data, this was adopted as a marine low reliability trigger value, for use only as an indicative interim working level.

Alcohol ethoxylates (AE)

A high reliability trigger value of 140 µg/L was derived for AE (normalised data) using the statistical distribution method with 95% protection. As there were limited marine data (which had similar or less sensitivity than freshwater species), this figure was adopted as a marine low reliability trigger value, for use only as an indicative interim working level.

The toxicities of cationic surfactants are generally greater than for the other surfactants (Lewis 1990, 1991).

References

ANZECC & ARMCANZ 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality, Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand, Canberra.

Dorn PB, Salanitro JP, Evans SH & Kravetz L 1993. Assessing the aquatic hazard of some branched and linear nonionic surfactants by biodegradation and toxicity. Environmental Toxicology and Chemistry 12, 1751-1762.

Feijtel TCJ & van de Plassche EJ 1995. Environmental risk characterisation of 4 major surfactants used in the Netherlands. Report 679101 025, National Institute of Public Health and Environmental Protection, Bilthoven, The Netherlands.

Guhl W & Gode P 1989. Correlations between lethal and chronic/biocenotic effect concentration of surfactants. Tenside Surfactants and Detergents 26, 282-287

Hennes-Morgan EC & de Oude NT 1994. Detergents. In Handbook of ecotoxicology, vol 2, ed P Calow, Blackwell Scientific, London.

Lewis MA 1990. Chronic toxicities of surfactants and detergent builders to algae: A review and risk assessment. Ecotoxicology and Environmental Safety 20, 123-140.

Lewis MA 1991. Chronic and sublethal toxicities of surfactants to aquatic animals: A review and risk assessment. Water Research 25, 101-113.

Kimerle RA 1989. Aquatic and terrestrial ecotoxicology of linear alkylbenzene sulfonate. Tenside Surfactants Detergents 26, 169-176.

Konemann WH 1981. Quantitative structure-activity relationships in fish toxicity studies. Part 1: Relationships for 50 industrial pollutants. Toxicology 19, 209-221.

OECD1992a. Report of the OECD workshop on extrapolation of laboratory aquatic toxicity data to the real environment. OECD Environment Monographs No 59, Organisation for Economic Co-operation and Development, Paris.

Nimrod AC & Benson WH 1996. Environmental estogenic effects of alkylphenol ethoxylates. Critical Reviews in Toxicology 26, 335-364.

Saarikoski J & Veluksela M 1982. Relationship between physicochemical properties of phenols and their toxicity and accumulation in fish. Ecotoxicology and Environmental Safety 6, 501-512.