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State of the Marine Environment Report for Australia: Pollution - Technical Annex 2

Edited by Leon P. Zann
David Sutton
Great Barrier Reef Marine Park Authority, Townsville Queensland

Ocean Rescue 2000 Program
Department of the Environment, Sport and Territories, Canberra, 1995

ISBN 0 642 17406 7

The problems of nutrients and eutrophication in the Australian marine environment

Jon Brodie
Great Barrier Reef Marine Park Authority
Townsville Qld


Eutrophication results from the supply of excessive plant nutrient substances to an aquatic ecosystem leading to enhanced plant growth or to a change in the composition of plant and other species. Coastal eutrophication is recognised as a worldwide and growing problem in areas affected by agricultural and urban run-off (GESAMP 1990; Nixon 1990; Smayda 1990; Rosenberg 1985). Some of the problems in Australian fresh and marine waters have been summarised in recent publications (e.g. Brodie et al. 1990; AEC 1987; Cullen 1986). The principal nutrients associated with eutrophication are nitrogen (N) and phosphorus (P) but others such as organic carbon, silicon, iron, molybdenum and manganese may play a supplementary role. On a global scale, it is now estimated that the input of nutrients to the oceans from human sources via rivers is equal to, or greater than, the natural input (Windom 1992).

The growth of aquatic plants is regulated by limiting biological, chemical and physical factors. Of the chemical factors, growth nutrients such as N or P may be present or available in the environment in limited amounts. The ratio of nutrient elements in the environment is also important, as plants need and use nutrients in a ratio linked to the composition of their own biomass. Thus it has been postulated that marine phytoplankton require carbon, N and P in the ratio 106:16:1 (Redfield 1934, 1958) - the Redfield Ratio - which corresponds to the elemental ratio of these elements in phytoplankton. However it is now clear that requirements vary among species though there are some common features in their requirements (Hecky & Killam 1988; Howarth 1988). The supply of silicon is also particularly significant in coastal phytoplankton dynamics, because of its importance as a structural component of diatoms. The ratio of silicon to P is often grossly changed by anthropogenic inputs to coastal areas, and this can lead to phytoplankton community shifts from diatoms to flagellates (Smayda 1990).

Seagrasses and marine macroalgae have different C:N:P ratios (Atkinson & Smith 1983) and changes in the supply of N or P may change conditions to favour the growth of one plant at the expense of others. Because of differences in the supply of a limiting nutrient, increased loading of that nutrient may produce very different responses in different ecosystems. In the case of Port Phillip Bay in Victoria it has been shown that the seagrass, Herterozostera tasmanica, has responded positively to N inputs because its growth was limited by N concentrations in the sediment porewater. In contrast, in Westernport Bay, where sediment porewater N concentrations are much higher, the seagrass showed little response to N enrichment (Bulthius et al. 1992).

The idea of nutrient limitation in aquatic environments has a long history (see for example Hecky & Killam 1988), particularly in freshwater lakes. Extensive research in recent decades, including artificial fertilisation of a set of lakes in Canada (Schindler 1975), has clarified the role of P limitation in lake ecosystems. This has led to an understanding of P loading (Vollenweider 1976) and an ability to make recommendations and predictions regarding the rehabilitation of eutrophic lakes. It was also clearly shown that lakes could eutrophy naturally and that anthropogenic influences can accelerate this process to decades rather than thousands of years. This idea of 'cultural' eutrophication enhancing an existing phenomenon has also been recognised in estuarine environments. Estuaries of south-west Australia are in a process of slow natural eutrophication and anthropogenic impacts may be accelerating this process dramatically (Hodgkin 1988).

In the marine environment the question of nutrient limitation of primary production is controversial (see Howarth 1988). While N has been commonly regarded as the limiting element in many marine ecosystems (Ryther & Dunstan 1971), P may be limiting in some systems (Smith 1984) and various combinations of simultaneous or alternating N or P limitation have been reported. In diatom populations, where silicon may also be a limiting nutrient, enhanced diatom growth due to anthropogenic inputs of this element is not likely as silicon is not a major component of wastewaters.

The principal nutrients involved in eutrophication in coastal waters, N and P, may exist in a range of forms or species. Inputs of N and P into the marine environment normally occur as a mixture of a number of chemical forms which vary considerably in many important properties relevant to eutrophication including bio-availability, mobility and stability. The principal forms of N include the dissolved inorganic species, nitrate, nitrite and ammonium/ammonia; dissolved organic N including simple, identifiable compounds such as amino acids and urea as well as complex, high molecular weight, poorly characterised material ; and particulate N. Additionally, dissolved nitrogenous gases may also play a role in marine N cycles: N gas is available in unlimited quantities from the atmosphere, and can be used by N-fixing organisms.

Phosphorus may occur as dissolved inorganic P (primarily orthophosphate but also as polymeric forms), dissolved organic P, particulate organic P, and particulate inorganic P. The latter may include, especially in river run-off, P mineral substances which are not available for uptake by marine organisms.

N and P are regarded as the major environmental nutrients and those we know most about. Carbon is generally regarded as nonlimiting. Other nutrient elements, especially micronutrients, have been less studied and little is known of environmental responses to their concentration or their distribution and availability in the marine environment.


Anthropogenic or anthropogenically-enhanced sources of nutrients to the marine environment are soil erosion, fertiliser run-off, sewage discharge, rainfall, intensive animal production including aquaculture, and industrial discharges.


The major anthropogenic inputs of P to the marine environment are agricultural run-off (diffuse inputs) and sewage discharges (point source inputs). The principal agricultural inputs derive primarily from the use of fertilisers and secondly from increased mobilisation of natural P from increased soil erosion. Sewage P sources are primarily human waste and P-containing detergents. In 1983 it was estimated that 5000 tonnes of P compounds were added directly to Australian coastal waters from marine sewage outfalls while coastal flowing rivers added 34 000 tonnes from diffuse sources and 1100 tonnes from sewage (Garman 1983). It has been estimated that in excess of 85% of all P entering Australian waters originates from diffuse sources (AEC 1987). However in low flow conditions in rivers, sewage may be the major source of nutrients. This has been clearly shown in studies on the sources of nutrients causing blue-green algal blooms in the Murray-Darling river system (GHD 1992), where sewage inputs predominate in dry conditions (when the blooms form) and diffuse inputs dominate during the wetter periods. Information on the relative magnitude of different sources of nutrients within catchments can be found in Rosich and Cullen (1982).

Phosphorus in soil is derived from weathering of phosphatic minerals, such as apatite, in the parent rock. Australian soils are, in general, extremely poor in P compared to soils found in North America and Europe (McLaughlin et al. 1992). It may thus be postulated that Australian coastal ecosystems have also developed in a low P environment. Phosphorus from soil inputs to the marine environment may not be completely available for uptake by marine plants and the phosphate buffer mechanism (Froelich 1988) exerts some control over available dissolved P concentrations.

An increase in use of P in fertilisers in Australia from 1951 is shown in Figure 1 (adapted from McLaughlin et al. 1992). In some catchments, similar or more dramatic increases in the use of fertilisers in this period have been documented. For example Figure 2 (adapted from Valentine 1988) shows the rise in total fertiliser use in Atherton Shire on the Barron River catchment (Qld) where, in offshore areas, potential problems with reef degradation due to enhanced nutrient loading have been recognised (Rasmussen & Cuff 1990). Figure 3 (adapted from Birch 1982) shows a similar rise in superphosphate usage in the Swan coastal plain catchments (WA), where major eutrophication problems have been experienced in the Peel-Harvey system. The sources of fertiliser P are almost entirely from imported P fertilisers or phosphate rock, the latter having been processed into fertilisers in Australia.

Figure 1: Consumption of P (expressed as elemental P) in Australia between 1951 and 1988. Adapted from McLaughlin et al. (1992)

Figure 1

Figure 2: Annual fertilizer use un Atherton Shire (based in ABS data: total of all types). Adapted from Valentine (1988)

Figure 2

Near major urban areas sewage input of P may exceed that from other sources. Moss et al. (1992) have shown that along the Queensland east coast, agricultural inputs of P (and N) are far greater than those from sewage except in south-east Queensland around the heavily urbanised areas of Brisbane, the Gold Coast and the Sunshine Coast. For the rest of the coast (Cape York to Fraser Island) the ratio of agricultural to sewage inputs is 30:1 while for the south-east the ratio is 0.6:1 (Moss et al. 1992).


The principal anthropogenic sources of N which may reach the coastal zone are also agricultural run-off and sewage discharges. Nitrogen fertiliser usage has been growing rapidly over the last three decades (Figure 4; adapted from McLaughlin et al. 1992). In 1987 approximately 371 000 tonnes were used in agriculture. The largest crop use of N was for wheat and sugarcane. Table 1 shows a breakdown of usage versus crop type (from Bellingham 1989).

Other significant agricultural sources of N include discharges from intensive animal industries such as piggeries and beef feedlots (see review by Bowmer and Laut 1992). Urban stormwater containing garden fertiliser, animal faeces, septic system leachate and sewerage system overflows may also be significant localised sources of N and other nutrients. Typical stormwater flows can contain concentrations of total N and P of the order of 2 mg/L-N and 1 mg/L-P (GHD 1981). A Victorian study (Weeks 1982) showed that the total flux of nutrients in stormwater from urban catchments may equal that in sewage discharges for the same population. Direct industrial discharges of wastewater may also contain considerable amounts of N (e.g., from a nickel refinery; Carey et al. 1982) but the discharge of such wastes has generally been reduced through compliance regulations in recent years.

Figure 3: Superphosphatefertiliser usage for coastal plain catchments adjacent to the Peel-Harvey system. Adapted from Birch (1982)

Figure 3

Figure 4: Consumption of N (expressed as elemental N) in Australia between 1951 and 1988 (Australian Manufacturers' Committee data). Adapted from McLaughlin et al. (1992)

Table 1: Estimates of nitrogen use (kt) in Australian agriculture in the 1987 crop year

Crop Qld NSW Vic Tas SA WA Total %
Sugar 55 3 0 0 0 0 58 15.6
Wheat 14 20 11 0 11 63 119 32.1
Barley, oats 6 5 5 0 8 14 38 10.2
Sorghum, maize, rice 19 16 0 0 0 3 38 10.2
Other crops 20 6 0 0 1 0 27 7.3
Fruit and vines 6 8 1 1 3 3 22 5.9
Vegetables 8 6 3 1 1 2 21 5.7
Pastures and fodders 17 6 9 1 3 8 44 11.9
Miscellaneous 1 2 1 0 0 0 4 1.1
Total 146 72 30 3 27 93 371  
% 39.3 19.4 8.1 0.8 7.3 25.1    

Source: Bellingham 1989

In catchments with a substantial industrial coverage, N can also be exported to coastal waters via N-enriched rainfall (Paerl 1985). This has been shown to be an important proportion of the N flux in some overseas coastal waters (Hinga et al. 1991; Paerl 1985) and may be significant in understanding nutrient limitation in some oceanic areas (Fanning 1989). There have been few published studies of such potential inputs in Australia except in the context of acid rain (AEC 1990) but around the major industrial cities of Newcastle, Sydney, Melbourne Wollongong and Geelong some input of this type may be significant. Inputs of N, from wash-out of N oxides to Port Phillip Bay, have been estimated to be of the order of 500 tonnes per year, about 10% of the total load (Carnovale & Saunders 1987). Other estuarine systems near our major cities could be expected to receive atmospheric loads of similar magnitudes.

Soil erosion

Estimates of increased soil erosion and hence increased inputs of natural soil nutrients are preliminary at present. On an Australia wide basis it is considered that in the arid zone, 880 000 km2 of the area used for grazing (3.36 million km2) is experiencing enhanced soil erosion (Woods 1983). For eastern Queensland catchments the estimates of Moss et al. (1992), which are derived from catchment modelling studies refined by results from a few well monitored catchments, show that approximately four times as much sediment, N and P now enter the marine environment than before the introduction of western agriculture. Much of the increased nutrient load is postulated to be natural soil nutrients mobilised by erosion from grazing (Figures 5 & 6; Moss et al. 1992). As there are considerable problems in the estimation of P losses from erosion using areal land-use figures (McLaughlin et al. 1992), the estimates of Moss et al. (1992) must be viewed as indicative.

Deforestation has been a recurring theme in Australia's development, the prime purpose being for agricultural development rather than forestry. Even in recent times clearing has continued, the most notable example being the extensive clearance of the brigalow (Acacia harpophylla) belt in Queensland (Bailey 1984). Several million hectares were cleared for cropping and grazing land in the Fitzroy River catchment from the mid 1960s till the early 1970s (Webb 1984). Deleterious effects on soil erosion and salinisation have been described by Webb (1984) and Johnson (1985). Effects extend downstream to the marine environment in the form of increased sediment and nutrient fluxes to the coastal zone (Byron et al. 1992; Moss et al. 1992). Table 2, adapted from Bucher and Saenger (1991), documents Australian catchments rated on the extent of clearance and shows the larger extent of clearance in the smaller south-eastern states with their greater population densities compared to Qld, WA and the NT.

Intensive logging practices, such as clear-felling, can produce major changes in river condition (including elevated concentrations of dissolved salts, suspended solids and nutrients) especially during peak flow periods (Campbell & Doeg 1989). The major downstream effect is increased sedimentation and this has been documented in the Eden area of south-eastern NSW (Olive & Rieger 1988). This study also highlighted the difficulty of detecting such changes in such a highly variable rainfall and stream flow regime.


Most sewage effluent discharged into coastal waters has been treated to a secondary level and as such contains most of the nutrients originally present in the raw sewage. Secondary effluent typically contains 30-60 mg/L N and 6-12 mg/L P. Other plant growth stimulating substances are commonly found in sewage effluents. Many secondary treatment plants achieve substantial reduction in N content in their effluent by aeration control. Nitrogen concentrations in the effluent may be reduced to below 15 mg/L in this way ; for example, the Coombabah plant on the Gold Coast, as described in Moss (1992). Approximately 10 000 tonnes of P and 100 000 tonnes of N are produced in sewerage effluent in Australia annually. As the majority of the Australian population lives near the coast much of this effluent will enter coastal waters.

In some cases effluent is reused for industrial water e.g. in the lower Hunter region (NSW) and in Kwinana (WA)(Hanrahan & Gale 1989); for irrigation on golf courses (e.g. Townsville and Cairns); for horticulture near Melbourne (Hanrahan & Gale 1989) as well as for agriculture on the Werribee Sewage Farm, Melbourne. Reuse for irrigation on golf courses and on gardens is mandatory on resort islands in the Great Barrier Reef (GBR) Marine Park (Brodie 1992). Reuse schemes have also been investigated in the Tuggerah Lakes (Wyong), Coffs Harbour and Lake Macquarie areas of NSW, but considerable problems in their application have prevented implementation (Hanrahan & Gale 1989). An investigation into large scale reuse for irrigation of plantation forests near Adelaide is under way (Anon 1991). Tertiary treatment plants which reduce nutrients are now being introduced in many inland situations but as yet only a few operate in coastal regions.


There are significant differences in the fate and behaviour of N and P in the marine environment. A number of micro-organisms have the ability to either fix atmospheric N or denitrify N compounds (principally nitrate) to N gas. Thus N can be derived from, or lost to, the enormous pool of atmospheric N. Phosphorus does not undergo such transformations but can be locked up in soils or sediments by adsorption and binding to the soil particles and become, in a practical sense, removed from the ecosystem. Nitrogen is less readily lost in this way.

Table 2: Distribution of catchment clearance a

% of Catchment Cleared of Natural Vegetation
State/Territory <25% 25-50% 50-75% >75% Insufficient Information Total
Queensland 170 17 4 3 113 307
New South Wales 20 31 25 5 - 81
Victoria 8 6 6 15 - 35
Tasmania 17 - - - 46 63
South Australia - 1 1 12 1 15
Western Australia 125 2 3 1 14 145
Northern Territory 136 1 - - - 137
Total 476 58 39 36 174 783

a Number of catchments in each category Source: Bucher and Saenger 1991

Particle-bound P is not easily leached by percolating water but can be mobilised by the movement of the soil itself - as when erosion occurs. Much of the P lost from agricultural lands is transported in particulate form. In the South Pine River in south-east Queensland, Cosser (1989) found during a 2 year study that over 77% of P was transported as particulates. He also noted that 80% of this P flux occurred during 2.8% (only 20 days) of the total time, during periods of intense rainfall and subsequent storm flow in the river. Thus control of soil erosion may also have major impacts on P transport to the coastal environment. It is estimated that the practice of green cane harvesting of sugarcane followed by the use of cane trash as a soil 'blanket' reduces annual soil erosion to 5 tonnes per hectare. This compares to traditional burning and tillage practices which yield average annual soil loses of about 150 tonnes per hectare. This method has a similar lowering effect on P losses (Prove & Hicks 1991). North of Townsville over 70% of cane is now harvested by the green cane method (Prove & Hicks 1991) and this is believed to have led to a significant decrease in P loading to the adjacent coastal seas (Kuhn 1990).

Sandy soils may behave somewhat differently than most other soils in that they can export a relatively high proportion of phosphate not bound to particulates. This was the situation found for the sandy, coastal plain catchments of the Peel-Harvey system (Hodgkin et al. 1985; Birch 1982).

The proportion of sediment-bound P which is available for uptake by marine organisms has been the subject of some debate but it is known that P can be desorbed from sediment particles in the mixing zone between freshwater and seawater, where salinity, pH and redox conditions change. A number of studies both in Australia (Carpenter & David Smith 1985) and overseas (Fox 1989) have demonstrated this process.

Figure 5: Comparison of magnitude of point and diffuse catchment sources of N. From Moss et al. (1992)

Figure 5

Figure 6: Comparison of magnitude of point and diffuse catchment sources of P. From Moss et al. (1992)

Figure 6

Nitrogen, on the other hand, is not as strongly bound to soil particles and is more readily lost to percolating water by leaching. Especially in the inorganic forms of nitrate and ammonium, N is very soluble in water and is easily moved by both overland and groundwater flows. Control of soil erosion is thus less effective in controlling N losses.

A significant factor in the transport of nutrients and sediment in river systems is the overwhelming contribution of the wet season (in climates with distinct wet and dry seasons), and storm flow. The problems in estimating fluxes under these conditions have been quantified in a number of overseas studies (Walling & Webb 1985) and is most difficult in semiarid climates with monsoonal rainfall regimes. Much of tropical Australia fits this latter category and it is recognised that Australian rivers have more variable annual river flows and annual floods than the rest of the world's continents (Finlayson & McMahon 1988). Major northern Australian rivers such as the Fitzroy (Brodie & Mitchell 1992), Burdekin, Tully and Johnstone (Arakel et al. 1989; Mitchell 1987) transport almost their entire flux of nutrients to the coastal zone during storm flow (normally cyclonic). Measurement of these fluxes is difficult as the episodes may only last a few days per year and access is limited.

Another natural source of N to coastal waters is N fixation by blue-green algae, particularly the pelagic Trichodesmium. Trichodesmium occurs widely in Australian tropical and subtropical waters (Creagh 1985) but its significance is unclear. Published N fixation rates vary 20-fold (Carpenter & Capone 1992) and the concentration of Trichodesmium in Australian waters is poorly known. It has been suggested that it is now more common in the GBR due to enhanced P inputs (Bell 1991) but there is limited evidence to support this.

Concentrations in the marine environment

The nutrient status of Australian marine waters is a function of hydrodynamics i.e. flushing times and the presence of adjacent sources of nutrient input e.g. currents, upwellings or sewage discharges. As a result of anthropogenic inputs most of the enclosed or semi-enclosed coastal water bodies in the southern half of the continent near major river estuaries, urban and industrial areas have elevated nutrient status and are eutrophic or showing signs of incipient eutrophication.

In general, oceanic waters surrounding Australian have low nutrient concentrations (Jeffrey et al. 1990). While upwelling systems occur in a number of areas they are limited and intermittent. The system on the western coast of Australia (Jeffrey et al. 1990) is far weaker than those characterising the continental western coasts of the Americas and Africa (Pickard & Emery 1990), being blocked by the southward-flowing Leeuwin current. In Bass Strait, nutrient concentrations are low throughout the year but show evidence of at least weak upwelling on the eastern boundary of nutrient rich sub-Antarctic water (Gibbs et al. 1986). Along the eastern coast of Tas. and Vic., and the coast of NSW, some evidence of intrusion of subsurface waters has been found. This is considered to be an important source of nutrients for these coastal areas (Rochford 1984). Episodic upwelling also occurs from the Coral Sea into the GBR lagoon. This probably forms an important nutrient supply for the GBR (Furnas & Mitchell 1986; Andrews & Gentian 1982).

For many water bodies the total flux of nutrients to the system may be a better indicator and predictor of eutrophication than nutrient concentrations. While the technique has been extensively used and refined for freshwater lakes, its use in the marine environment is often much more difficult to apply due to the difficulty in quantifying natural fluxes.

Australia is a continent of low surface run-off to its coastal seas and has no large river systems in terms of discharge. The larger rivers in terms of catchment size and length (the Coopers Creek/Thomson and the Warburton Creek/Diamantina systems) either do not discharge to the sea at all, or have a low discharge (the Murray-Darling). By comparison the Fly River, which partially discharges into Australian coastal waters from Papua New Guinea, yields a volume of water and sediment similar to that of all Australian rivers combined (see Figure 7; from Harris 1991). Thus Australian coastal regions are not naturally exposed to continuous high levels of terrestrial run-off. Australia also has few large coastal embayments or seas (e.g. Gulf of Carpentaria) comparable to Chesapeake Bay, Puget Sound, the Red Sea, the Baltic Sea or the Gulf of Thailand, where the discharge of a large river system might be retained.

Figure 7: Sediments discharge of rivers entering the Great Barrier Reef Lagoon and location of terrigenous shelf and Holocene coastal sediments. From Harris (1991)

Figure 7

Table 3: Nutrient concentrations at which observable increases in plant growth have occurred in various water bodiesa

Waterbody/Area Total Nitrogen (g/L) Total Phosphorus (g/L)
Hawkesbury Nepean 650 55
Peel/Harvey 150 25
Lake Burley Griffin 90b 60
Lake Macquarie 600 60
Murray River 550 40
Kosciusko National Park 360 40

a It must be stressed that these are approximate thresholds only. These waterbodies show a gradation of effects both with time and location and hence it is not possible to be definitive as to exactly where and when eutrophication is apparent.

b Only oxidised forms of nitrogen were measured.

Source: AEC 1987

Gulf of Thailand, where the discharge of a large river system might be retained.

Much of our knowledge of the general nutrient status of Australian coastal waters has come from the work of the CSIRO and recently, in northern waters, from the Australian Institute of Marine Science (AIMS).

Substantial data sets describe concentrations of N, P and silicon species in coastal and shelf waters; for example Bass Strait (Gibbs et al. 1986 ), the east coast (Rochford 1984), the GBR lagoon (Furnas 1991; Andrews 1983; Revelante & Gilmartin 1982) and the North West Shelf (Mackey 1984; Rochford 1977).

The concentrations of N and/or P which indicate, or indeed cause, eutrophication have been long debated. Obviously these critical concentrations depend on the ecosystem in question, with some able to better cope with nutrient stress than others. Table 3 from the Australian Environment Council's review of nutrients in Australian waters (AEC 1987) lists the nutrient concentrations at which 'observable increases in plant growth have occurred in various water bodies' (both fresh and marine). Of note is the similarity in the critical concentrations between the diverse water bodies, from saline coastal lakes and estuaries to large rivers and inland artificial impoundments. The report concludes that 'visual evidence of eutrophication is likely to occur if total N concentration is within or exceeds the range of 400-600 mg/L and/or total P concentration is within or exceeds the range of 40-60 mg/L'.

Environmental impacts

Generally eutrophication in coastal ecosystems is derived from increased concentrations of N and P enhancing plant growth. In any locality the species which best respond to enhanced nutrients depends on the interaction between physiological limitation factors and competition between plant types. Thus while in isolation, coral zooxanthellae, seagrasses, phytoplankton and benthic algae may all respond positively to an increased nutrient supply, in mixed communities the response of one may dominate. While low levels of nutrient enhancement may promote seagrass growth (Lukatelich et al. 1987), at higher levels epiphytic algal overgrowth occurs, reducing light and leading to seagrass demise. This has been reported in Cockburn Sound, WA (Cambridge & McComb 1984) and Gulf St Vincent , SA (Neverauskas 1987). In some situations secondary effects may dominate the benthic response, such as filter feeders (barnacles, tube worms, sponges) expanding their coverage in response to increased phytoplankton concentrations. However, light limitation due to high turbidity, may inhibit plant growth even though nutrient levels are high, as noted in the Brisbane River (Moss 1987) and as suggested to occur in the Georges River, NSW (Heath et al 1980) and Lake Bonney, SA.

Eutrophication will often progress through a sequence of stages, characterised in the global State of the Marine Environment Report as an idealised progression involving: '(a) enhanced primary productivity, (b) changes in plant species composition, (c) very dense blooms, often toxic, (d) anoxic conditions, (e) adverse effects on fish and invertebrates, (f) impact on amenity, (g) changes in structure of benthic communities' (GESAMP 1990). Not all these stages will always be present or evident.

Overseas, coastal eutrophication has been most evident in enclosed and semi-enclosed seas and estuaries. The most prominent examples include the northern Adriatic Sea where annual noxious algal blooms now occur (Justic 1987); the Baltic Sea with anoxia problems (Larsson et al. 1985); the North Sea with increasingly regular algal blooms including the toxic species Chrysochromulina polylepis (Underdal et al. 1989) and the nuisance alga Phaeocystis pouchetti (Lancelot et al. 1987); the Black Sea with increasing anoxia problems, loss of fisheries and blooms of introduced species (Mee 1992); the Inland Sea of Japan (Seto Sea) (Yasui & Kobayaski 1991); and Chesapeake Bay with loss of benthic fauna and fisheries (Officer et al. 1984).

In addition to large nutrient fluxes, the principal factors implicated in eutrophication in many of these overseas localities is long water residence times (i.e. poor flushing) promoted by the enclosed nature of the water body. This has the effect of allowing build-up of nutrients during periods of high input. Many Australian coastal water bodies of similar long residence times and high nutrient inputs are showing analogous eutrophication.

Coastal lagoons represent 11.4% of the Australian coastline (Cromwell 1971, cited in King & Hodgson 1986). In all areas affected by urban and agricultural run-off, these lagoons and lakes are showing stages of incipient eutrophication as a result of human activity. This is generally true for that part of the southern mainland Australian coastline stretching from Perth to Brisbane. Prominent examples include the Peel-Harvey system, Gippsland Lakes and many of the NSW coastal lakes. In addition, many of the embayments and estuaries in this region are affected, including Cockburn Sound, Gulf St Vincent, Westernport Bay, Botany Bay and Morton Bay. For the few very large systems in Australia which could be threatened by eutrophication (the GBR lagoon and the Gulf of Carpentaria) convincing evidence of effects is not available, although claims have been made that a problem exists in the GBR (Bell 1992).

Eutrophication in enclosed coastal waters is comparable to that commonly seen in recent years in Australia's inland water bodies, particularly the Murray-Darling system (GHD 1992). This has led to blue-green algal blooms (frequently toxic) with consequential effects on human and stock drinking water supplies and in-river ecological effects. The causes of the eutrophication (sewage and agricultural run-off), are the same as for coastal regions with the added problem in rivers of a lack of water flow at the end of the dry season, which leads to enhanced nutrient concentrations.

Loss of seagrass beds

One of the most common features in the record of effects of Australian coastal eutrophication has been a loss of seagrass beds. Shepherd et al. (1989) reviewed twelve well described cases of major seagrass loss following enhanced nutrient or sediment supply. Walker and McComb (Table 1; 1993), ascribed eutrophication as a cause of the losses in most of the cases.

In Cockburn Sound (WA), there has been increasing urban and industrial development of the adjacent coast from 1954, and increased inputs of industrial discharges and sewage during the 1960s (Cambridge et al. 1986), led to a 97% loss (3300 hectares) of seagrass beds by 1978 (Figure 8; Cambridge & McComb 1984). The loss has been attributed to overgrowth by epiphytic algae (Cambridge et al. 1986; Silberstein et al. 1986) leading to substantial reductions in ambient light reaching leaf surfaces. The decline in the seagrass coverage was closely correlated with increasing N loadings to the Sound. Recent reductions in nutrient loadings have occurred, including a new sewage outfall outside the Sound (1984) and better quality industrial discharges (1982). These management actions have led to a halt in the decline of the seagrass beds but only minimal recovery and regrowth (Hillman 1986).

In Gulf St Vincent (SA), the principal Adelaide sewage outfalls at Glenelg and Bolivar, and the sewage sludge outfall at Semaphore have caused substantial losses of seagrass in their vicinity (Figure 9; Shepherd et al. 1989; Neverauskas 1987). The total loss of seagrass area is estimated to be over 5000 hectares. As in Cockburn Sound, the principal cause of the loss has been epiphytic overgrowth leading to light loss. This has been accompanied by shifts in species composition (Neverauskas 1987).

Figure 8: Seagrass loss in Cockburn Sound. Each map shows Coockburn sound surrounded by the coast of the mainland to the right, and Garden Island to the left. The 10m contour line is indicated. The shading shows the area of seagrass meadow present at different times. From Cambridge and McComb (1984)

Figure 8

Westernport (Vic.) has suffered almost complete destruction of the extensive seagrass beds present before 1973 (Coleman 1982; Bulthius 1981). By 1984 the area of macrobenthic plants in the bay had declined from the 25 000 hectares of 1973 to 7200 hectares (Bulthius, unpublished, in Shepherd et al. 1989). Remaining areas of seagrass have also declined in biomass, with an 85% decrease in the standing crop between 1975 and 1984. The causal factors are not clearly identified but general development in the catchment leading to increased sedimentation, turbidity and nutrient levels in the Bay has been implicated. The principal cause appears to have been increased sedimentation combined with the stress of higher temperatures on the most severely affected intertidal beds (Bulthius 1983a, 1983b). The loss of seagrass appears to be closely correlated with an 80% reduction in the catch of King George Whiting.

In NSW a large number of bays, lagoons and river estuaries have suffered loss of seagrass areas in recent years. Anthropogenic factors have been implicated in many cases but as King and Hodgson (1986) note, the natural high variability in seagrass area and growth makes monitoring and detecting anthropogenic changes difficult. Changes in techniques and methodologies over the years have made comparisons difficult.

In the coastal embayments of NSW, substantial changes in the area of seagrass beds have been documented in Lake Illawarra (King 1988a; Yassini 1985); Botany Bay (Larkum & West 1990); Tuggerah Lakes (King & Hodgson 1986; King & Holland 1986); Lake Macquarie (Simmons & Trengrove 1989; King & Hodgson 1986) and the estuaries of the Georges, Clarence and Tweed Rivers (Shepherd et al. 1989). Some doubt as to the extent of losses due to anthropogenic causes versus natural change exists for Lake Illawarra (King 1988a) and the causes of loss in Botany Bay are also not clear (McGuinness 1988). For Lake Macquarie and the Tuggerah Lakes, catchment urbanisation (particularly stormwater run-off and point source discharges), combined with agricultural run-off have led to increased nutrient loadings. As a result, increases in phytoplankton, turbidity and macroalgae have occurred, along with decreased clarity and losses of seagrass areas. The estimated losses in seagrass areas are 700 hectares for Lake Macquarie and 1300 hectares for the Tuggerah Lakes (Shepherd et al. 1989). Studies of the Clarence River estuary have shown that the present seagrass beds (Zostera spp.) only occur over 20% of the area described in the 1940s (NSW Government 1992).

Figure 9: Time course of seagrass loss from Barker Inlet to Gawler River from 1954 to 1985. From Shepherd et al. (1989)

Figure 9

Other systems showing loss of seagrass possibly correlated with terrestrial run-off are Princess Royal Harbour and Oyster Harbour near Albany, WA (Bastyan 1986); Port Lincoln, SA (Shepherd 1975); Peel-Harvey Inlet, WA (discussed elsewhere in this review) and possibly Morton Bay, Qld (Kirkman 1978). In northern Australia cyclonic freshwater inundation may lead to massive seagrass loss but subsequent recovery e.g. in Cleveland Bay in 1971 associated with Cyclone Althea (Pringle 1989) and Hervey Bay associated with Cyclone Fran in 1992 (Preen 1993). The losses are probably associated with prolonged freshwater stress and loss of light due to turbidity. In the recent event in Hervey Bay, over 1000 km2 of seagrass have been lost resulting in significant mortality and migration of the dugong population and reduction in commercial prawn and fish catches. The loss/recovery cycle in tropical Australia may be contrasted to southern Australia, where seagrass loss from eutrophication may be more gradual but recovery has not been recorded. After physical damage to Posidonia beds recovery is also slow as shown in the lack of regrowth in areas of seagrass in Jervis Bay, denuded in the 1960s by military bombing practice (Beckmann 1991).

Low levels of nutrient enhancement may favour the growth of seagrass beds. For example, at Wilson Inlet (WA), where enhanced seagrass growth has been documented, catchment nutrient loads were elevated above natural loads, but were much lower than in the Peel-Harvey system, and macroalgae were less prominent (Lukatelich et al. 1987). At Green Island (north Qld) the sewage discharge from a primary treatment plant was trapped by prevailing wind and tidal conditions in a hydrodynamic retention area to the north of the island. The extensive expansion of the seagrass beds which occurred (Van Woesik 1989) has been generally attributed to the effects of the outfall, although it may also have been promoted by changes in sediment composition due to nearby channel dredging.

Nutrients may also enhance mangrove growth. In areas where increasing siltation has provided 'new' habitat for mangroves at the expense of seagrass beds, extra nutrients can lead to vigorous mangrove growth. This has occurred in a number of estuaries in the Sydney area (Dunstan 1990) including Botany Bay (Georges River), Brisbane Waters, and the Lane Cove (McLaughlin 1987) and Cooks Rivers. At high levels of nutrient enrichment the mangroves may be affected by excessive algal growth. This has been the case in Barker Inlet and Chapman Creek near Adelaide, where discharges from the Bolivar sewage treatment plant has caused increased growth of the green alga Ulva. Drifts of the alga directly shade or smother newly established mangrove seedlings. Seedling mortality up to ten times levels in unimpacted areas has been observed (Edyvane 1991).

Algal blooms and red tides

The most visible indicator of coastal eutrophication is extensive blooms of phytoplankton and/or benthic macroalgae. In many overseas localities eutrophication in large water bodies has been characterised by blooms of planktonic species such as Phaeocystis (the North Sea; Davidson & Marchant 1992; Lancelot et al. 1987) and Noctiluca miliaris (the Black Sea; Mee 1992), and mucus aggregates (the northern Adriatic Sea; Stachowitsch et al. 1990). Some of the algae are toxic and may cause fish kills while others are aesthetically 'nuisance' algae, causing spoiling of beaches, offensive odours and slimy water. Many cause secondary blooms of undesirable fauna, such as the ctenophore Mnemiopsis leidyi (the Black Sea; Mee 1992) which has reached bloom biomass densities of 1 kg m-2.

In Australia a large number of estuaries, bays and coastal lakes have begun to experience algal blooms in the last thirty years . The best known examples of phytoplankton and macroalgal blooms come from the Peel-Harvey system in WA, while the effects of toxic algae are best known from Tasmania and Victoria, where the closure of shellfish beds has resulted from blooms of an algae suspected to have been introduced into Australia in ballast water (Hallegraeff & Bolch 1992).

The Peel-Harvey system provides the best example of an Australian marine eutrophic system and the range of impacts caused. Located south of Perth in WA, it consists of two connected coastal lagoons which form the estuary of three small rivers: the Murray, the Harvey and the Serpentine (Figure 10). The catchment has been extensively modified for agricultural uses, mainly beef, sheep and dairy production (Yeates et al. 1984), and extensive use of phosphatic fertilisers has occurred over the last forty years. Phosphorus input, in particular, has risen by factors of nine times for the Serpentine River and fifty times for the Murray River (Hodgkin et al. 1981) in the period between 1949-56 and 1972-78.

Figure 10: Peel-Harvey Estuary. From McComb and Lukatelich (1986)

Figure 10

The two lagoons show distinctly different eutrophic responses. In Peel Inlet the main response has been extensive growth of the macroalga Cladophora in the early years of the problem, changing more recently to Chaetomorpha and Ulva as dominant species (Lavery et al. 1991). The Harvey Estuary, in contrast, has little growth of macroalgae but massive blooms of the blue-green alga Nodularia spumigena (McComb et al. 1981). The macroalgae in the Peel Inlet break free from the lagoon floor, accumulate on the beaches and rot, producing nauseating odours. They have been mechanically harvested offshore and regularly removed from beaches, but this has only been partially successful in keeping beaches near populated areas free from the rotting algae (Hillman et al. 1990). In the Harvey estuary the growth of macroalgae is inhibited by light attenuation due to high water turbidity. This seems to be due to the geographic orientation of the estuary (Figure 10), which lies with its long axis parallel to the prevailing winds and hence has high wind-resuspension rates (Gabrielson & Lukatelich 1985). As a result, while not having macroalgal problems, the Harvey estuary experiences massive blooms of Nodularia spumigena in late spring and early summer. With the exception of 1979 and 1987, the blooms have occurred each year since 1978 (Hillman et al. 1990) and are closely correlated with the freshwater inflow from the Harvey River, with its associated P load (McComb & Lukatelich 1986). Figure 11 shows the relationship between Nodularia chlorophyll a and P load for the period 1977 to 1983 (modified from McComb & Lukatelich, 1986). Nodularia also has a nauseating smell, especially when it accumulates and rots on the shores, and has been blamed for sickness in local residents (Hodgkin et al. 1985).

Dense blooms of Nodularia spumigena have affected fish and crab populations in the Harvey estuary. This has been shown by reduced commercial fish catches and reduced fish numbers in areas with high chlorophyll-a levels, indicative of high Nodularia density (Lenanton et al. 1985; Potter et al. 1983). However increased macroalgal biomass in the Peel has led to an increase in commercial fish catch (Lenanton et al. 1984).

Figure 11: Relationship between maximum average concentration of chlorophyll a in HarveyEstuary in summer, and the amount of phosphorus entering the estuary in winter from the Harvey River and associated drains. From McComb & Lukatelich (1986)

Figure 11

The south-east coast of Australia, from Port Phillip Bay to Fraser Island, has many coastal lagoons, lakes and bays similar in a geomorphological sense to the Peel-Harvey system. They have narrow connections to the sea, are fed by rivers flowing from small to moderate sized catchments and are shallow with low tidal influences and long water residency. They vary in salinity regime, size, riverine inflow, rainfall and flushing period and any wetlands associated with them reflect these physical differences (Bucher & Saenger 1991). Those water bodies that have documented problems with algal blooms are listed in Table 4, and their location shown in Figure 12. From these it is clear that the problem is widespread, especially near centres of high population and agricultural development.

The largest of the relatively enclosed systems is Port Phillip Bay. At various times over the last two decades concern has been expressed that the Bay was becoming eutrophic. Monitoring and research studies have attempted to resolve the issue. Those of the 1970s established only minor changes to water quality and benthic communities (Axelrad et al. 1981; Brown et al. 1980; EPA 1979; Poore & Rainer 1979). However continued concern about the effects on the Bay of urban development in the catchments and the long-term options for disposal of treated sewage effluents has led to further studies and monitoring by a variety of organisations, particularly the Victorian EPA (e.g. Lukatelich 1990) and Melbourne Water (e.g. Bremner et al. 1989), culminating in a recently commissioned large-scale study coordinated by the CSIRO (1992).

In NSW nuisance algal blooms have been recorded over a number of years in many coastal water bodies. Cheng (1981) lists Tuggerah Lakes, Narrabeen Lagoon, Dee Why Lagoon, Harbord Lagoon, Avoca Lake and Lake Illawarra as having major blooms while also noting that lagoons with less populated and developed catchments such as the Myall Lakes and Smith Lake remain relatively unspoiled.

In the Tuggerah Lakes the nuisance algae are mostly green macroalga of the genera Enteromorpha, Chaetomorpha and Rhizoclonium (King 1988b). The lake area is 80 km2, the catchment area 670 km2 while only a 1% exchange of water with the ocean occurs on each tidal cycle under normal entrance conditions (Anon 1990). The catchment is significantly urbanised and falls entirely within the Wyong Shire Council area. There has been a 7-8% annual population growth in the Shire between 1973 and 1990 (Anon 1990). Considerable areas within the catchment are used for citrus and other crops while a large coal fired power station (Munmorah - 1400 MW) uses lake water for cooling purposes. Potential problems in the lake system were recognised as early as 1969 (Higginson 1970). Controversy exists over whether hot water discharges from the power stations or nutrient enhancement are responsible for algal blooms, with the evidence favouring nutrient enhancement (King 1988b). As in the Peel-Harvey system the excessive growth, accumulation, and subsequent decay of algae, producing offensive odours, is the major focus of concern.

Table 4: Australian coastal areas showing eutrophication

Locality Effects
Swan River Estuary Phytoplankton blooms
Peel-Harvey Estuary Phytoplankton, macroalgal blooms
Cockburn Sound Seagrass loss
Wilson Inlet Minor enhanced seagrass growth
Albany Harbour Seagrass loss
Port Lincoln Seagrass loss
Gulf St Vincent Major seagrass loss, toxic algae
Port Phillip Bay Macrophyte growth, toxic algae
Western Port Major seagrass loss
Gippsland Lakes Phytoplankton blooms
Derwent Estuary Phytoplankton blooms, toxic algae
Huon Estuary Phytoplankton blooms
Lake Illawarra Seagrass loss, macrophyte growth
Botany Bay Seagrass loss
Avoca Lagoon Phytoplankton blooms
Harbord Lagoon Phytoplankton blooms
Tuggerah Lakes Seagrass loss, macrophyte growth
Lake Macquarie Phytoplankton blooms
Clarence Estuary Seagrass loss
Tweed Estuary Minor seagrass loss
Moreton Bay Phytoplankton blooms
Great Barrier Reef Macrophyte growth, coral in Lagoon decline

Considerable changes in the floristic composition of Lake Illawarra appears to have taken place since the early 1960s, in which time a number of studies of the lake occurred (King 1988a; King & Barclay 1986; West 1985; Evans & Gibbs 1981; Harris et al. 1980; Harris 1976; Higginson 1968). King (1988a) noted that this temporal variation is common in lakes and drew attention to the problems of detecting anthropogenically-induced changes against an unknown natural background variation. However Lake Illawarra is regarded as having undergone significant eutrophication, manifested mainly in the form of excessive filamentous algal growth (LIMC 1986).

Other Australian coastal systems to have experienced blooms of either benthic macroalgae or phytoplankton include Morton Bay where red tides were followed by fish kills (Moss 1987), Sydney Harbour (Revelante & Gilmartin 1978), Gippsland Lakes (Poore 1982), the Hawkesbury River and estuary, the Derwent River estuary (Coleman 1983), the Huon River estuary, the Swan River estuary (John 1987; Hodgkin & Vicker 1987 in John 1987), the Tweed Estuary (Anon 1986) and Orielton Lagoon (Buttermore 1977). Other systems such as Lake Macquarie have experienced a slow decline in water clarity with increasing nutrient concentrations (SPCC 1983), the latter attributed to rising nutrient input from an increasingly urbanised catchment (Simmons & Trengrove 1989). Even the oceanographically dynamic seas off Sydney have very recently (January 1993) experienced widespread red algal blooms extending from Wollongong to the Hawkesbury estuary. Whether the recent relocation of the principal Sydney ocean sewage outfalls offshore into deeper waters will contribute to changes in the phytoplankton community remains to be seen.

Toxic red tides

An increasing problem in coastal waters worldwide is the occurrence of toxic phytoplankton blooms (Smayda 1990). These blooms, intensified by eutrophic coastal conditions, cause finfish kills and shellfish to become toxic. In Tasmanian waters, blooms of the dinoflagellate Gymnodinium catenatum have occurred since 1986 (Hallegraeff et al. 1989), and caused temporary closures of commercial shellfish beds. From distributional evidence and the absence of the organism in sediment records, it has been suggested that G. catenatum was introduced into Tas. via ships' ballast water (Hallegraeff & Bolch 1992). In SA the toxic dinoflagellate Alexandrium minutum now blooms annually in the Port River, near Adelaide (Cannon 1990; Hallegraeff et al. 1988), causing toxicity in shellfish (Oshima et al. 1989). This organism is also thought to be introduced. Toxicity in wild mussels has resulted from blooms of Alexandrium catenella in Port Phillip Bay, but effects on shellfish farms have been minor (Hallegraeff et al. 1991). The incidence of harmful algal blooms in the Australian region has recently been reviewed by Hallegraeff (1993a).

The increasing incidence of 'toxic red tides' around the Australian coast has led to claims that these will now be a regular event and that their principal cause is the discharge of sewage effluents into the ocean (Illert 1993). While algal research workers discount the more tenuous links of the 'ecological holocaust' claim, they acknowledge the continuing discharge of poorly treated sewage effluents into the oceans is undesirable (Hallegraeff 1993b).

Long-term effects on benthos near sewage outfalls

Ocean sewage outfalls are best located in deep water, on open, hydrodynamically active coasts to maximise dilution and dispersion. In Australia many large outfalls have been relocated to such positions in recent years, having previously discharged into rivers, coastal lakes and bays or onto the ocean shoreline. Examples include Devonport, Tas. (Wallis & Holmes 1987); Cape Peron, WA (Chalmer & Edmonds 1986); Geelong, Vic. (McLearie & Barkley 1987); Latrobe Valley, Vic. (Sampson & Howard 1987); Sydney (North Head, Bondi and Malabar; Fagan et al. 1993); Lake Macquarie and Tuggerah Lakes (Norah Head) in NSW, and eastern Melbourne (Cape Schanck) (Brown et al. 1990). Problems of eutrophication from well-placed shoreline ocean outfalls on active coasts, some of which have been discharging for many decades, appear to have been very localised in the examples which have been studied.

Monitoring of the Cape Schanck outfall from before the commencement of discharge (1975) until 1988 has indicated that effects on intertidal macroalgal communities were restricted to within one kilometre of the outfall (Brown et al. 1990; Manning 1979). Studies of algal flora and intertidal invertebrates adjacent to the Sydney shoreline outfalls (now moved offshore) also revealed very localised effects, with green algal mats predominating at distances of up to 0.5 km from the outfall, but little measurable effect at distances over one kilometre (Fairweather 1990; May 1985).

The Great Barrier Reef

The GBR is the only Australian marine ecosystem which is comparable in size to areas overseas where large scale eutrophication has occurred (such as the Black and Baltic Seas), and which has been suggested as becoming eutrophic (Bell 1992, 1991). Bell's evidence is drawn mainly from comparison of phytoplankton records from near Low Isles (Qld) in 1928/29 (Marshall 1933) with others collected more recently off Townsville in the late 1970s and early 1980s (Revelante & Gilmartin 1982; Walker & O'Donnell 1981). While the conclusions have a limited scientific and statistical weight, the claims have promoted debate as to the possible eutrophication of the GBR (Bell & Gabric 1991; Kinsey 1991a; Walker 1991).

Queensland coastal catchments have been extensively modified since European settlement by forestry, urbanisation and agriculture, particularly sugarcane cultivation and beef grazing. A recent report (Moss et al. 1992), using catchment modelling calibrated with known run-off data, has estimated that 15 million tonnes of sediment, 77 thousand tonnes of N and 11 thousand tonnes of P are exported to the Queensland eastern coastal zone via river discharges. This is approximately four times the load estimated for the pre-European settlement period. The load in the GBR region is mostly from agricultural run-off, with sewage a secondary and minor contribution. The relative increase in nutrient load in recent times is comparable to that calculated for river inputs to the Black and North Seas, where eutrophication has occurred. Coral reefs, which normally exist in very low nutrient conditions (Kinsey 1991b), are particularly sensitive to nutrients and the algal and filter-feeder overgrowth which occurs in conditions of elevated nutrient supply.

Evidence that a eutrophication problem already exists on the GBR is patchy. Anecdotal evidence from local residents and long-term regular visitors suggests that many reefs were in 'better' condition in past decades than at present. This is supported to a limited extent by the sparse historical photographic records of reefs in locations such as Magnetic Island, Low Isles and the Whitsunday Islands. The claim in most of these cases is that the reefs in question are now far more dominated by algae than in the past. More compelling scientific evidence comes from coral cores taken from reefs offshore from Cairns. The corals from which the cores were taken range up to more than a century old. The yearly growth bands, which provide a record of the environmental conditions at the time the band was growing, have been interpreted as suggesting that significant changes in the growth of coral in this area began to occur about fifty years ago. This correlates with the introduction of the intensive agricultural use of fertilisers in the Barron River catchment on the adjacent coast (Rasmussen & Cuff 1990). The available evidence from nutrient and plankton sampling in the GBR lagoon has not indicated regional or temporal increases in nutrient concentrations (Furnas 1991) but it is acknowledged that systematic monitoring data is scant (Brodie & Furnas 1992).

The difficulty of determining whether the GBR is becoming eutrophic in the face of a limited historical data record and naturally variable environmental conditions reflects similar difficulties experienced in other Australian systems, such as Botany Bay (McGuinness 1988) and NSW coastal lakes (King 1988a). The development of scientific tools to clearly separate natural from anthropogenic changes, and to provide high levels of certainty about historical environmental conditions is a priority for the measurement of long-term degradation.


In Australia, monitoring of nutrient concentrations, nutrient enrichment and eutrophication can be conveniently divided into a few broad categories depending on time and space scales implicit in the objectives of each monitoring initiative. These are: long-term monitoring to derive ambient conditions and detect trends; monitoring associated with known or suspected eutrophic systems; monitoring associated with local developments and discharges; and monitoring associated with ecological process studies. Australian marine monitoring programs have been briefly considered by Williams and Gilmour (1986).

Long-term monitoring

Few long-term records of nutrient or phytoplankton concentrations, taken on a regular basis at consistent sites in Australian coastal waters, are available. For example, Table 5 shows a number of oceanographic stations sampled by CSIRO over long periods. The logistical problems of collecting such data are formidable, many of the samples in this data set being sampled by lighthouse keepers and island research station staff. Most of the stations are no longer sampled and at present there is no national coastal water quality monitoring program. Generally this important issue is one that 'slips through the cracks' of agency responsibility in the Commonwealth and State sectors. Currently the issue of long-term monitoring is under discussion by ANZECC.

Table 5: CSIRO Coastal Station Network - coastal water quality monitoringa

Station (position) Attempted frequency (per month) Depth range (m) Parameters b Date Commenced Date Finished
Heron Island 1 - 2 0 - 50 T,S,NO3,SiO4 Dec 1976 1989
Norfolk Island 1 - 2 0 - 50 T,S,NO3,SiO4 Dec 1977 1989
Lord Howe Island 1 - 2 0 - 50 T,S,NO3,SiO4 April 1976 1989

Port Hacking

1 - 4 0 - 50 T,S,NO3,SiO4
some DO,PO4
1942 Active
Port Hacking
1 - 4 0 - 100 T,S,NO3,SiO4 1953 Active
Maria Island 1 0 - 50 T,S,NO3,SiO4 1944 Active
Rottnest Island 1 - 2 0 - 50 T,S,NO3,SiO4 1970 Active
Booby Island 1 - 2 0 - 10 T,S,NO3,SiO4 June 1977 1984
Lizard Island 1 - 2 0 - 25 T,S,NO3,SiO4 Aug 1974 1984
Geraldton 1 - 2 0 - 40 T,S,NO3,SiO4 Dec 1978 1985
Low Isles 1 - 2 0 - 10 T,S,NO3,SiO4 June 1977 July 1982
Port Macdonell 1 0 - 50 T,S,NO3,SiO4
some DO
1973 1981
Eden 1 - 2 0 - 50 T,S,NO3,SiO4 1974 1986

a CSIRO have long term records of basic hydrological parameters for Australian coastal waters, approximately 100 000 data points. The data is generally recorded at 6 depths.

b Abbreviations: T - Temperature, S - Salinity, NO3 - Nitrate, SiO4 - Silicate, PO4 - Phosphate, DO - Dissolved Oxygen.

Source: CSIRO, Division of Oceanography, Tasmania

Some long-term records of water quality or benthic conditions have been kept by individuals or small research groups, but these are site specific and the information gathered has not always been published. A good example is the unpublished thirty year record of benthic cover at fixed quadrats collected by Connel at Heron Island . Another data set, collected by the Queensland Department of Environment and Heritage over a 14 year period from 1979 to the present, consists of monthly chlorophyll concentrations from four stations stretching from inside the Southport Broadwater (Qld) to 5 km out to sea (Moss 1992). Another long-term monitoring data set, from Port Phillip Bay, also includes chlorophyll a data (Brown 1988).

The use of remote sensing for monitoring the nutrient status of coastal waters has been investigated. Satellite sensors measuring chlorophyll by colour has been the method of choice (e.g. Gabric et al. 1990), using various platforms such as the Coastal Zone Colour Scanner (CZCS - now defunct), Advanced Very High Resolution Radiometer (AVHRR), Landsat and SPOT. Although CZCS worked well in oceanic and clear water conditions, turbidity and coloured humic material severely limited its use in coastal waters. The other platforms still operating also have serious limitations (Muller-Karger 1992) such as poor spatial resolution (AVHRR) or high cost and poor spectral resolution (Landsat and SPOT). Aircraft sensors can also be used and CSIRO is pioneering this technique in Australia. Aircraft platforms can avoid the problems of cloud interference experienced with satellite sensors but may also be very expensive to operate.

Monitoring eutrophic and potentially eutrophic systems

Many of the coastal systems known to be, or suspected of being, either in an advanced state of eutrophication or becoming eutrophic, have had their status monitored. Examples include the Peel-Harvey system, eastern Gulf St Vincent, Cockburn Sound, Lake Illawarra, Lake Macquarie and the Tuggerah Lakes. State environment authorities often have programs at a regional scale for coastal water bodies. For example the Victorian EPA operates long-term monitoring in Port Phillip Bay, Westernport Bay and the Gippsland Lakes. Major new programs commenced in recent years include an intensive research and monitoring program in Port Phillip Bay (CSIRO 1992) and a program incorporating monitoring of fish, benthos, water quality and sediment quality in the GBR region (Brodie & Furnas 1992; P. Moran, AIMS, pers. comm.).

Development, discharge and compliance monitoring

Monitoring programs, usually imposed as a condition of approval by the responsible environmental authority, are often associated with developments such as marinas and harbours, or discharges such as sewage effluent, industrial wastewaters and power station cooling waters. Examples of this type of monitoring include the CSIRO monitoring in Jervis Bay as part of the environmental impact assessment (EIA) for a naval development (Ward & Jacoby 1992); monitoring of the ocean environment off Sydney after the installation of deep-water, offshore sewage outfalls (Fagan et al. 1993); monitoring of thermal water discharges from power stations discharging into NSW coastal lakes (e.g. King & Hodgson 1986); monitoring of the effects of building and operating a marina adjacent to a coral reef (Brodie et al. 1992); and monitoring a nickel refinery wastewater outfall discharging ammonia-rich wastes (Carey et al. 1982).

Many of the monitoring programs of this type have a very limited scope and the results are not widely disseminated or part of the public domain. In addition most environmental management authorities do not require such monitoring to be carried out by independent researchers, the work often being carried out by the developer or discharger, or consultants reporting directly to them. Thus the results obtained, and their interpretation, are often open to criticism of bias or distortion. On the other hand many such programs have resulted in data sets generating valuable ecological information. Examples would include the work of Brown et al. (1990) on the long-term variation in algal intertidal flora both close to and some distance from the Cape Schanck sewage discharge, and studies investigating long-term changes in benthic flora in NSW coastal lakes receiving cooling water discharges from coal-fired power stations (King & Hodgson 1986).

Ecological and oceanographic monitoring

A large proportion of the published studies which could be described as monitoring have been associated with studies of ecological processes. Monitoring is an integral part of the information required to understand temporal and spatial patterns in populations of biota, geochemical fluxes and oceanographic forcing in ecosystems.

Two organisations which have been heavily involved in biological and chemical oceanographic studies are CSIRO, some of their work already having been described in this report (e.g. Mackey 1984; Rochford 1984), and the AIMS. AIMS has worked principally in the GBR region and their biological oceanography group has published records of nutrient concentrations and nutrient processes over the last 15 years (e.g. Furnas 1991; Andrews 1983).

Many marine ecological studies of nutrient/ecosystem interactions included a monitoring component and these have been carried out at many locations around Australia. A few examples include an examination of long-term variations in subtidal algal floras (Jervis Bay and Ulladulla; May 1985, 1981) and macrobenthos ( Hawkesbury Estuary; Jones, 1987); comparative studies of communities in different parts of the coast (e.g. zooplankton in Port Phillip Bay and Westernport Bay; Kimmerer & McKinnon 1985) and of macrobenthic fauna of seagrass beds in a number of bays in NSW (Collett et al. 1984); and ecosystem studies in the extensive seagrass beds of Shark Bay, WA (Walker 1989).

Information Systems

Given the lack of any national water quality monitoring programs or national coordination of the monitoring which does occur (note this is now being addressed under the aegis of the National Water Quality Management Strategy), it is not surprising that there is no central database of Australian monitoring results. The Environmental Resources Information Network (ERIN; a subunit of the Commonwealth Department of the Arts, Sport, Environment and Territories), is in the process of setting up a National Marine Information System (NMIS or NATMIS) which should eventually fulfil this need. The CSIRO, through its Coastal and Marine Resource Information System (CAMRIS), is also attempting to coordinate the storage of its national data holdings and further develop applications as part of the CSIRO Coastal Zone Program, as well as link with ERIN and the National Resource Information Centre (of the Department of Primary Industries and Energy) initiatives for database access.

Another activity which may improve the coordination of knowledge of the state of coastal areas is the estuary inventory of Bucher and Saenger (1991, 1989). This inventory 'was compiled to gain a national perspective on the ecological status of estuaries and to identify research and management priorities'. Similar efforts at a State level have previously been made, such as the NSW estuarine inventory of West et al. (1985).


The complexity in processes and interactions of coastal ecosystems makes modelling an attractive option. Models can be used to integrate information on physical, chemical and biological processes to assist understanding of the system and to predict spatial and temporal change in parameters of interest to environmental management. Models may be numerical, physical or prototype (NSW Government 1992). They are widely used to estimate pollutant loadings (e.g. Moss et al. 1992), to simulate ecosystem response to increased nutrient loadings ( e.g. the Coastal Ocean Ecology Model used in WA; Van Senden & Button 1992), and to provide managers with tools for scenario-building and assessing the implementation of environmental management strategies (e.g. the Port Phillip Bay study; CSIRO 1992). The use of estuarine models in Australia was reviewed by Beer (1983) who considered examples from Westernport Bay, Blackwood River, Peel Inlet, Port Hacking and Gippsland Lakes. Shortcomings in their use were noted but the success of the Peel Inlet example highlighted.


Management of the marine environment varies among states. Systems such as 'assimilative capacity' and 'ecosystem approach' have been published as an overall philosophy in some states (Pearce 1991), but such unifying ideas tend to change with time, circumstances and governments. The current inquiry by the Resource Assessment Commission into the management of the coastal zone in Australia may lead to a greater unity of approach.

Management of nutrient loading to coastal environments, and any subsequent problems of eutrophication, is based on the reduction of nutrient-rich effluents from the land or better dispersion of existing discharges. This may be accomplished by prevention of nutrient generation processes such as control of soil erosion, changes in the use or nature of fertilisers, reuse rather than discharge of nutrient rich effluents, diversion of discharges into less sensitive or better flushed environments, engineering works to improve flushing, and nutrient removal from effluents. Rehabilitation of existing eutrophic systems has also been attempted in some places, using methods such as replanting of seagrasses (Hillman 1986) and removal of nutrient-rich sediments. Some Australian examples of management to ameliorate eutrophication in coastal systems are listed below.

Engineering and land management


To manage and remedy the previously-described problems of eutrophication in the Peel-Harvey system, a three-fold strategy is being implemented (Gorham et al. 1988) involving:

  1. catchment management using controls over fertiliser usage, increased use of slow release fertilisers and controls over land clearance;
  2. continuation of mechanical algal clearance;
  3. cutting a channel from the Harvey Estuary to the ocean (the Dawesville Channel) to improve flushing in the system (Figure 12).

The introduction of slow-release fertilisers, particularly New Coastal Superphosphate, is expected to have a major impact on the amount of P leached from the Peel-Harvey catchment (Yeates et al. 1984). New Coastal Superphosphate is a mixture of superphosphate, rock phosphate (the slow release component) and sulphur and supplies adequate P and sulphur to previously fertilised soils without excessive leaching. Controls over water movement by building locks, tree planting and creation of small wetlands on drainage lines is also being used to slow the movement of drainage water to the estuaries (Cribb 1993).

The Dawesville Channel will be 1.5 kilometres long, involve the excavation of 4.5 million cubic metres of material and cost $56 million. Tidal flushing combined with reduced residence time for the estuarine water is expected to reduce nutrient buildup and decrease the incidence of algal blooms. The channel will also increase tidal ranges in the estuaries and shift the estuarine conditions towards a more marine state, a change causing some concern in the area (Cribb 1993). The channel is planned to be completed by 1996.

Tuggerah Lakes

The restoration program for the Tuggerah Lakes also depends on a combination of rehabilitation strategies. The three identified objectives are to reduce sediment and nutrient inputs to the lakes, to remove sediment and nutrients from the lakes and to improve tidal exchange (Anon 1990). These objectives are to be achieved by:

  1. construction of sediment traps and nutrient filters on streams and drains discharging into the lakes;
  2. clearing aquatic plant accumulations from the lake beaches;
  3. removing silt and aquatic plants from the inshore lake bed;
  4. deepening channels to improve water circulation and navigational access;
  5. implementing measures to improve tidal flushing of the lakes. (The principal measure will involve some form of entrance deepening);
  6. catchment management measures including sewerage upgrades, erosion control, stormwater treatment, stream bank stabilisation and wetland preservation.

The restoration program is planned to take four years at an approximate cost of $10 million from NSW Government funds. This is in addition to spending some $3.2 million on lake rehabilitation over the period from 1980 to 1988, funded from a variety of sources, but in particular the Wyong Shire Council (Anon 1990).

Catchment management

The principal tool now being used in Australia for land use management, of relevance to nutrient run-off into the coastal zone, is the various forms of catchment management (known in NSW as Total Catchment Management or in Queensland and Victoria as Integrated Catchment Management). These schemes bring together local landholders, state and local government organisations and other interested parties in an attempt to manage catchments as a whole. Some of the methods relevant to nutrient run-off and used within catchments as components of management strategies include minimum tillage and stubble retention agricultural systems, revegetation of stream banks, buffer strips along stream banks, measures to prevent erosion along roadways and during road construction, sewage system upgrades, preservation of wetlands as sediment and nutrient 'filters', controlling stocking rates, contour cultivation; better fertiliser management including timing of use, subsoil injection and the use of slow-release types; and minimising erosion during urban development.

Sewage nutrient discharge minimisation

A number of strategies have been suggested for the minimisation of nutrient inputs to coastal areas from sewage discharge, and most of them are being implemented or trialed somewhere in Australia. They can be grouped into minimisation of nutrients entering the sewage system, better dilution/dispersion methods, reuse of effluents, and nutrient reduction before discharge.

Phosphate from detergents is a major component of the P content of sewage, estimates suggesting it may contribute 65% of the load in domestic effluent (Nicholls et al. 1977). A ban on phosphates in detergents was a major plank of the USEPA's policy to reduce nutrient loading to the US Great Lakes in the 1970s, and is believed to have been relatively successful. It has been suggested that similar bans may be desirable in the GBR region (Chiswell & Hammock 1991; Bell 1989). Recently released reports of the task force examining the blue-green algal problems of the Murray-Darling system have also contained similar recommendations. However reduction in P loading will not be as effective in N-limited systems as it is in P-limited lake and river systems. The replacement of phosphate in detergents may also lead to other coastal marine problems if the replacement is a N compound (Ryther & Dunstan 1971) or zeolite/polycarboxylic acids which have caused slime formation in the Adriatic Sea (MacKenzie 1993).

As previously mentioned, the relocation of sewage outfalls into waters with better dilution and dispersion characteristics has been a common management response to sewage problems in recent years. The schemes usually involve moving outfalls from discharging into rivers, lakes or shorelines to discharge into deep offshore conditions. Typical examples are the Latrobe Valley scheme (Sampson & Howard 1987), Lake Macquarie and Tuggerah Lakes schemes and the Sydney offshore outfall program (Fagan et al. 1993).

Tertiary treatment of sewage to reduce N and P loads in the effluent is not commonly practiced in Australia, but with the outbreak of blue-green algal problems in Australian river systems such measures will now be implemented in many inland sewage plants. Some coastal facilities are already running nutrient reduction methods. Examples include Port Macquarie, where N stripping occurs (AEC 1987) and a number of island resorts in the Great Barrier Reef Marine Park which have N stripping (Radisson Long Island), P stripping (Daydream Island) or both (Green and Lindeman Islands). Most extended aeration sewage treatment plants can be operated to reduce N levels to less than 15 mg L-1 and biological P removal is also possible in such plants. A review of the use of nutrient reduction in Australian sewage plants and the methodology available can be found in the Australian Environment Council publication 'Nutrients in Australian Waters' (AEC 1987).

The increasing practice of reuse of effluents for land irrigation has already been discussed in this report. Other schemes for land use of sewage include Werribee near Melbourne where land irrigation of sewage has been practiced since 1897. A final effluent equivalent to secondary treatment is discharged into Port Phillip Bay (Bremner & Chiffings 1991). The scheme has been successful with only minor undesirable effects observed in Port Phillip Bay (Axelrad et al. 1981), but an intensive study of the Bay has now commenced to check on its environmental condition in relation to inputs from Werribee, the Yarra River and other catchment sources.

Reuse of effluent for residential nonpotable application is growing overseas but is not permitted in Australia under current regulatory practices (Wilkins & Anderson 1991). A pilot project to assess operating requirements, risk and community acceptance was run at Shoalhaven Heads (NSW) from 1989 to 1991, the water being mainly used on gardens. The engineers operating the scheme concluded that such reuse can be made viable without compromising public health (Wilkins & Anderson 1991).

The National Health and Medical Research Council and the Australian Water Resources Council have now developed guidelines for the use of reclaimed water in Australia (Anon 1987).


Some research on the re-establishment of seagrass beds in Cockburn Sound has been done and is continuing (Hillman 1986; Nelson, pers. comm.). Results have been disappointing so far, with sediment instability a major factor contributing to failure to revegetate. Methods using artificial seagrass mats are now being investigated.


The effects of eutrophication have been estimated to cost Australia from $10-50 million per annum (Garman 1983). While it is difficult to make estimates of such costs, especially indirect ones from secondary effects, some published cost elements are described below. These appear to be in excess of Garman's estimates, perhaps reflecting the increasing costs involved since those estimates were made.

Structural measures required to reduce the effects of algal growth are estimated to have cost $170m in Cockburn Sound and $50m in the Peel-Harvey system (AWRC 1992). The smell and appearance of a persistent toxic bloom in the summer of 1987/88 in the Gippsland Lakes area is estimated to have cost $6.5m in lost tourist revenue (AWRC 1992). Economic loss due to reduction in fisheries catches are more difficult to quantify. In some instances catches have been claimed to have been enhanced while losses have been reported in other nearby areas e.g. enhancement in the Peel Inlet versus losses in the Harvey Estuary (Lennaton et al. 1985; Lennaton et al. 1984).

The costs of preventative measures for marine pollution may be very large. In Sydney the renovation of the inadequate sewage system is being partially financed through a household levy (The Special Environment Levy) of $80/year, which is planned to raise $485 million over 5 years. The total Sydney Clean Waterways Programme is budgeted for $7.1 billion over 20 years.

Summary and comments

Eutrophication in the Australian marine environment is characterised by those factors which make the Australian situation unique as well as those features common to the increasing worldwide incidence of coastal eutrophication. Australia has few rivers which have sizeable discharge and few large semi-enclosed coastal water bodies. However it does have a very large number of smaller enclosed and semi-enclosed coastal lakes, lagoons and bays, and a population highly concentrated (almost 80%) in the coastal zone. Most of our coastal catchments have been extensively modified by agriculture, forestry and urbanisation and the introduction of intensive agriculture in them has been a feature of recent decades. Australian soils are nutrient poor and nutrient run-off to the coastal zone was relatively low before European modification of the catchments. There are few large scale ocean upwelling systems on the coast and in combination with the low run-off, Australian coastal waters are naturally nutrient poor and relatively unproductive. They are thus particularly susceptible to nutrient pollution. The evidence presented in this review documents that almost all coastal water bodies of long residence times in the settled part of the Australian coast (from Cairns south around the coast to Spencer Gulf and the south-west of WA) have some effects of enhanced eutrophication (Table 4). With increasing urbanisation of the coast (notably north and south of Perth, the northern and southern NSW coast, the Sunshine and Gold Coasts and Hervey Bay, and north of Cairns) and further agricultural development on the coastal catchments, the problems will continue. Management by way of Integrated Catchment Management, sewage effluent reuse and system rehabilitation, while already initiated in some places, must continue on a national scale. A nationally coordinated approach to monitoring and management, as is being pursued through initiatives such as the National Water Quality Management Strategy, will assist in this and benefit Australia.

Acknowledgments: Many people contributed material to this review or assisted with comments on the manuscript. I would like to acknowledge their help, especially Karen Edyvane, Jock Brodie, Arthur McComb, Stella Humphries, Andrew Moss, Chris Crossland, Jan Forbes, the GBRMPA librarians, Helene Marsh and Gilianne Brodie.


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