Oils and petroleum hydrocarbons in freshwater and marine water
Toxicant default guideline values for protecting aquatic ecosystems
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
Crude oils are composed mainly of aliphatic and aromatic hydrocarbons with small amounts of sulfur-containing compounds such as thiophenes and thiolanes, and other more polar compounds (Volkman et al. 1994). The chemical and physical characteristics of oil determine its fate in the environment and its environmental effects. These properties, however, vary greatly. Tissot and Welte (1984: cited in Volkman et al. 1994) analysed 500 world oils and defined a typical oil as comprising 57% aliphatic hydrocarbons, 29% aromatic hydrocarbons and 14% asphaltenes and polar compounds containing nitrogen, sulfur and oxygen. Most crude oils produced in Australia are classified as light crudes, suitable for production of petrol, diesel and aviation fuels (Volkman et al. 1994). Heavy oils are imported from the Middle East, Indonesia and elsewhere.
Sources in the environment
Around 3.2 million tonnes of oils and petroleum hydrocarbons enter the marine environment in an average year. Volkman et al. (1994) estimated the sources to be oil spills from tankers (22%) and other transport (13%), natural oil seeps and biological processes (8%), urban and industrial sources (30%), bilge flushing, oil terminals and refineries (3%), offshore production (2%), and atmosphere fallout (9%). Other estimates (AIP & APPEA 1996) have been: industrial discharge and urban runoff, 37%; vessel operation, 30%; tanker accidents, 12%; atmospheric fallout, 9%; natural sources 7%; and exploration and production 2%.
Inputs to freshwater have often been due to accidents with road or rail tankers or exploration/production on land (Green & Trett 1989).
Oil is less dense than water and is biodegradable. As it floats on the surface of water, a major effect of oil on the environment results from shoreline smothering, unless it is first physically or chemically dispersed. In confined environments (e.g. small freshwater streams or lakes) biodegradation will result in reduction in dissolved oxygen while there can be a localised build-up of toxic fractions (Green & Trett 1989).
When oil is spilt at sea, the rate of weathering depends on the nature of the oil, water temperature, wave action and use of dispersants. Initial weathering processes depend on spreading of the oil, evaporation, dispersion, formation of emulsions, dissolving of oil and oxidation. After a few days, sedimentation and biodegradation take over as the main removal processes (ITOPF 1986).
As oil is not a single homogeneous product it is not possible to be prescriptive about its toxicology or to derive guideline figures using the standard procedure. Many studies have focussed on the effects of oil in the field, sensitivity of whole ecosystems and time of recovery. This is an appropriate approach, and is reflected in the information reported in the coastal resource atlases for oil spills (e.g. NSW EPA 1994). It is difficult in laboratory studies to mimic the exposure of organisms to oil and oil products in the field (Chapman et al. 1990).
The most exact values for oil toxicity studies are those from water-soluble fractions (WSF) (Anderson et al. 1974), where the oil concentrations are measured. Even then, continuous exposure (Singer et al. 1990) may not adequately reflect the changing conditions of a marine or estuarine environment (Pace et al. 1995).
The most toxic fractions of oil generally appear to be the lighter fractions, often containing higher proportions of aromatics. These include petroleum and diesel, although the higher volatility of petroleum limits exposure of organisms.
The information cited below is from Green and Trett (1989) except where otherwise indicated. Lower concentrations of oil can stimulate algal growth and enhance biodegradation processes. Oil spills in flowing water have less toxic effect on vegetation than those in standing water.
Crude oil WSF was toxic (LC50) to Daphnia magna at between 750 mg/L (72 hours) and 2110 mg/L (24 hours) in natural lake water. Shaken cultures of Norman Wells crude were toxic (LC100) to rotifers above 10 µg/L (nominal). The measured toxicity of Dubai crude WSF to the crustacean, Asellus aquaticus, was 11.6 mg/L (48-hour LC50). Norman Wells crude, both fresh and weathered, caused significant reproductive impairment to Daphnia pulex at 1 mg/L (Wong et al. 1981). Some indication of the toxicity of refined petroleum can be obtained from proportions of low molecular weight aromatic compounds. Toxicity of No. 2 diesel fuel oil to D. magna increased with water soluble fraction (WSF) concentration and with temperature: 48-hour LC50s at 10°C were 88% WSF and at 25°C, 10% WSF. Invertebrate mortalities in a number of tests were sometimes attributed to physical entrapment by the oil but there were notable differences in toxicities of different oils.
The toxicity of oils to freshwater fish varies widely with different oils and the method of exposure during the test. Oil may also cause tainting of fish flesh and loss of their invertebrate and plant food supply. The 24-hour LC50 of a mixed crude to Pimephales promelas was around 12 µL/L under flow-through conditions (Hedtke & Puglisi 1982) but much lower toxicity has been reported for different species under static conditions.
Toxicity of oils to freshwater species, such as D. magna and Gammarus lacustris, appears to increase with use of dispersants. For refined products, petrol was more toxic to shad Alosa sapidissima (48-hour LC50 of 91 mg/L) than diesel (167 mg/L) or Bunker C fuel (2417 mg/L) (Tagatz 1961). Most freshwater studies indicated that coal oil and shale oil were significantly more toxic than petroleum-derived oil.
Spillages of diesel into creeks have often resulted in kills of invertebrates, even if fish survive (Green & Trett 1989), and the invertebrates are slow to recover. In a spill in Alaska, diesel measuring 8 mg/L was associated with a 90% decrease in invertebrate numbers and loss of four taxa (Green & Trett 1989). This can result in indirect effects such as loss of food sources for fish and an increase in algal growth (Chapman & Simmons 1990). Oil can persist for longer if it is entrapped in sediments (Guiney et al. 1987).
The data on toxicity of oils to marine organisms is more extensive.
Gilbert (1996) reported company data that the toxicity of 14 crude and refined oils to fish and invertebrates was between 1 and 100 mg/L (48 to 96-hour LC50). Much of the research effort has focussed on the effects of oil spills on shoreline habitats such as mangroves, saltmarshes, seagrasses and mudflats. This information is used to develop priority rankings for shoreline protection and cleanup, as reported in coastal resource atlases (e.g. NSW EPA 1994).
Direct toxicity effects have been noted in the field following diesel spills (McEnally & Thompson 1989), particularly for fish, benthic organisms, bivalves, copepods, crabs and other invertebrate animals. A severe spill in Massachusetts in 1969 significantly depleted saltmarsh grasses for up to 12 years, and crabs for up to 7 years (National Research Council 1985).
Tsvetnenko (1998) has collated and screened toxicity data for six crude oils, including four from the NorthWest Shelf of Australia, two gas condensates, one diesel fuel and two bunker fuels. Only tests conducted at temperatures > 15°C and on water-soluble fractions prepared by standard methods were considered. Toxicity ranges from this review (normalised to account for degradation, as per Hamoda et al. 1989) are provided in Table 8.3.24.
Toxicity of dispersed oil is often greater than the oil and dispersant alone, although not always so. The use of dispersants can, however, prevent or minimise adverse effects in sensitive shoreline environments and can minimise penetration of oil into sediments (National Research Council 1989).
Table 8.3.24 Toxicity ranges for different oils to marine organisms—from Tsvetnenko (1998). LC50 figures normalised (Hamoda et al. 1989) (mg/L)
|Organism||Crude oil (n=6)||Gas condensate (n=2)||Diesel (n=1)||Bunker fuels (n=2)|
|Crustaceans (n=8)||0.07a-7.8||0.5-0.6||0.3-> 4.5||0.6-1.0|
|Algae (n=6)||0.5-> 10||10.6-11.5||0.5-> 1.6||NA|
a. 0.7-0.36 were reported for NW Shelf crudes to Penaeus monodon; the lowest figure for other crude oils to crustaceans was 0.2 mg/L to the crab Ocypode quadrata; NA = not available; note that the no. of species and no. of oils is a maximum no. and not all species were tested against all oils.
Tsvetnenko (1998) used the USEPA methods (Stephan et al. 1985, USEPA 1994d) to derive a final chronic value of 7 µg/L total petroleum hydrocarbons (TPH).
Previous guidelines for TPH (USEPA 1986, Train 1979) have recommended that for protection of aquatic life, the TPH concentration should not exceed 0.01 of the lowest continuous-flow 96-hour LC50 to several important freshwater and marine species. Given the limited amount of flow-through toxicity data, the normalisation procedure of Tsvetnenko (1998), based on Hamoda et al. 1989, would give equivalent data.
The use of the 100-fold factor is consistent with the assessment factor approach in these guidelines.
The ranges that are cited in Table 8.3.24 (from Tsvetnenko 1998) could be used if there were no data on the specific oil in question. The trigger value (low reliability) can be calculated by applying an assessment factor (AF) of 100 to the lowest acute figure in the appropriate range.
AIP & APPEA 1996. Petroleum topics: Oceans and oil spills. Australian Institute of Petroleum Ltd & Australian Petroleum Production & Exploration Association Ltd, November 1996.
Anderson JW, Neff JM, Cox BA, Tatem HE & Hightower GM 1974. Characterisation of dispersions and water-soluble fractions of crude and refined oils and their toxicity to crustaceans and fish. Marine Biology 27, 75-88.
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.
Chapman JC, Johnston NAL, Nelson PF, Sunderam RM & Thompson GB 1990. Acute toxicity of the water-soluble fraction of coal tar to the tiger prawn, Penaeus monodon. Oil and Chemical Pollution 6, 279-287.
Chapman JC & Simmons BS 1990. The effects of sewage on alpine streams in Kosciusko National Park. Environmental Monitoring and Assessment 14, 275-295.
Gilbert TD 1996. Proceedings of the Sixth National Plan Scientific Support Coordinators Workshop, Tasmania, December 2–6, Australian Marine Science Association, Belconnen, ACT.
Green J & Trett MW 1989. The fate and effects of oil in freshwater. Elsevier Applied Science, London.
Guiney PD, Sykora JL & Keleti G 1987. Environmental impact of an aviation kerosene spill on stream water quality in Cumbria County, Pennsylvania. Environmental Toxicology and Chemistry 6, 977-988.
Hamoda MF, Hamam SE & Shaban HI 1989. Volatisation of crude oil from saline water. Oil and Chemical Pollution 5, 311-331.
Hedtke SF & Puglisi FA 1982. Short-term toxicity of five oils to four freshwater species. Archives of Environmental Contamination and Toxicology 11, 425–430.
ITOPF 1986. Fate of marine oil spills. International Tanker Owners Pollution Federation Ltd, Technical information paper 11, London.
McEnally JM & Thompson GB 1989. Coastal resource atlas for oil spills in Jervis Bay. State Pollution Control Commission, Sydney, ISBN 0 7305 5288 8, 32pp.
National Research Council 1985. Oil in the sea: Inputs, fate and effects, National Academy Press, Washington DC.
National Research Council 1989. Using oil spill dispersants on the sea. Committee for Effectiveness of Oil Spill Dispersants, NRC, Washington DC.
Pace CB, Clark JR & Bragin GE 1995. Comparing crude oil toxicity under standard and environmentally realistic exposures. In Proceedings, 1995 International Oil Spill Conference, Long Beach, California, American Petroleum Institute, Washington DC, 1003-1004.
Singer MM, Smalheer DI & Tjeerdema RS 1990. A simple continuous-flow toxicity test system for microscopic life stages of aquatic organisms. Water Research 24, 899-903.
Stephan CE, Mount DI, Hansen DJ, Gentile HJ, Chapman GA & Brungs WA 1985. Guidelines for deriving numerical national water quality criteria for the protection of aquatic organisms and their uses. US Environmental Protection Agency, Washington DC. PB85-227049, 9811.
Tagatz ME 1961. Reduced oxygen tolerance and toxicity of petroleum products to juvenile American shad. Chesapeake Science 2, 65-71.
Train RE 1979. Quality criteria for water. Castle House Publications Ltd, Billing & Sons, Guilford, London and Worcester.
Tsvetnenko Y 1998. Derivation of Australian tropical marine water quality criteria for protection of aquatic life from adverse effects of petroleum hydrocarbons. Environmental Toxicology and Water Quality 13, 273-284.
USEPA 1986. Quality criteria for water. US Department of Commerce, National Technical Information Service, US Environmental Protection Agency, Springfield, Virginia. PB87-226759, EPA 440/5 86-001.
USEPA 1994d. Whole effluent toxicity (WET) control policy. US Environmental Protection Agency Office of Water, Washington DC. EPA/833-B-94-002.
Volkman JK, Miller GJ, Revill AT & Connell DW 1994. Oil Spills. In Environmental implications of offshore oil and gas development in Australia: The findings of an independent scientific review, eds JM Swan, JM Neff & PC Young. Australian Petroleum Exploration Association, Sydney, 509-695.
Wong CK, Engelhardt FR & Strickler JR 1981. Survival and fecundity of Daphnia pulex on exposure to particulate oil. Bulletin of Environmental Contamination and Toxicology 26, 606-612.