Agricultural development and urbanisation in the coastal areas has lead to a marked increase in nutrient loads entering estuarine and coastal waters. Nationally, nutrients entering marine waters through river discharge account for about 85% of nutrients in coastal zone waters (Brodie 1997). Land clearing, grazing and the use of agricultural fertilisers are recognised as being the primary causes of increased catchment nutrients and domestic sewage and industrial effluents can be significant near major urban centres (Brodie 1997; Koop and Hutchings 1997b). This has resulted in eutrophic conditions and increased algal blooms in many estuaries and bays around Australia (Brodie 1997).


Australian marine waters are naturally nutrient poor due to a combination of factors including the poor nutrient status of Australian soils, the small quantity of freshwater runoff, and the absence of major upwelling systems. Nutrient deficiency in turn limits primary production. Eutrophication alters the ecosystem via the following biological progression (GESAMP 1990):

  • Increased primary production;

  • Changes in plant species composition;

  • Very dense, often toxic, blooms;

  • Conditions of hypoxia (low oxygen concentrations) or anoxia (no oxygen);

  • Adverse effects on fish and invertebrates; and

  • Changes in structure of benthic communities.

Eutrophic waters typically support a high standing stock of attached algae or phytoplankton (Brodie 1997), the latter also producing an increase in turbidity resulting in decreased light penetration that affects benthic plants and corals.


Adverse effects on marine invertebrates can thus range from massive kills caused by the low levels of dissolved oxygen or the toxins released from dense blooms of noxious dinoflagellates, to more indirect effects resulting from habitat deterioration or shifts in community structure to favour more pollution tolerant taxa. For instance, the turbidity caused by increased growth of microscopic phytoplankton can have serious consequences for the growth and survival of seagrass beds and coral reefs, both of which rely on adequate transmitted light and therefore, relatively clear waters (Engel and Thayer 1998; Section 5.3.2).


Seagrass die-off due to eutrophication, with associated losses of fauna, has been documented in several locations (e.g., Hutchings et al. 1991; Section 6.4.2). Most light reduction is caused by increased water turbidity and an increase in the biomass of epiphytes on the seagrass leaves (Larkum 1976; Hutchings et al. 1991; State of the Environment Advisory Council 1996; Butler and Jernakoff 1999; Longstaff et al. 1999). In Princess Royal Harbour, Albany (south-western Australia), Peterson et al. (1994) documented dramatic declines in the abundance of two species of suspension-feeding bivalves (Katelysia scalarina and K. rhytiphora) – previously dominant components of the fauna – from around 160 m-2 in 1983-1985 to nearly zero in 1992. In addition to the crash in adult abundances, recruitment was negligible compared to that observed in 1983-85 (Peterson et al. 1994). These declines co-occurred with eutrophication, seagrass die-off and macroalgal blooms. Although the mechanisms for the decline are unknown, it was seen as the result of degradation in the ecosystem due to water quality decline (Peterson et al. 1994). Increasing nutrients can also lead to large increases in the area of seagrass beds - for example around Green Island on the GBR where untreated sewage is discharged from the island (Woesik 1989; cited in Brodie 1997).


The effects of increased nutrients on coral reefs, which occur in oligotrophic waters, are now fairly well known (e.g., Scott and Cope 1990; Kinsey 1991b; Morton 1994; Brodie 1997; Tomascik et al. 1997; Causey in press) but the precise ways in which reefs respond to these increases is poorly understood (Brown and Howard 1985; Hatcher et al. 1989; Grigg and Dollar 1990; McCook et al. 1997; Koop et al. 2001). Only a few studies were based on existing sewage discharges on the reef (e.g., Smith et al. 1981; Grigg 1995) or actual eutrophication and pollution gradients (Tomascik and Sander 1985; Tomascik and Sanders 1987a; Tomascik and Sanders 1987b; Tomascik et al. 1997) to demonstrate changes. Most studies, however, have been confined to laboratory experiments, which give limited insights into how entire reefs respond to elevated nutrients (e.g., Hoegh-Guldberg and Smith 1989; Hunte and Wittenberg 1992; Hoegh-Guldberg 1994; Yellowlees et al. 1994).


Increased growth of algae and phytoplankton are stimulated by nutrients such as nitrogen and phosphorus. The phytoplankton can result in decreased light penetration affecting the deeper-water corals and (together with increased bacterioplankton) can also encourage filter-feeding organisms that compete for space with coral, smother it, or that result in increased bioerosion by filter-feeding organisms. For example, Pari et al. (1998) showed that the intensity of bioerosion by grazing increases dramatically when reefs are exposed to pollution. Holmes (1997) examined sponges along a eutrophication gradient in coral reefs in Barbados, and found boring clionid sponges near the most eutrophic site (41%) compared to the least eutrophic site (24%). Furthermore, the abundance of branching corals was positively related to the frequency of boring sponges, suggesting that increased bioerosion may be partly responsible for community shifts toward branching corals in polluted waters (Holmes 1997). While several studies (e.g., Wilkinson 1987; Wilkinson and Cheshire 1990 {Caribbean and Great Barrier Reef}; Meesters et al. 1991 {Curaçao and Bonaire}; Zea 1994 {Colombia}) have found evidence suggesting that sponge communities react positively to nutrient enrichment, a couple (e.g., Muricy 1989; Schroeter et al. 1993) have reported decreases, possibly due to increased levels of sedimentation blocking canals and tissues.


Elevated phosphorus concentrations can also reduce calcification and hence the density of the coral skeleton making the colony more brittle and susceptible to damage (Kinsey and Davies 1979; Rasmussen and Cuff 1990). In addition, increased phytoplankton resulting from increased nutrients may possibly result in increased survival of crown-of-thorns starfish larvae.


There are grave concerns about increasing nutrient levels on the Great Barrier Reef (e.g., Bennell 1979; Bell 1991; Kinsey 1991a; Bell and Elmetri 1995; Wachenfeld et al. 1998; Koop et al. 2001) since with only 17% of catchments adjacent to the GBR are now considered to be in a natural condition ( cited in Koop et al. 2001; Gilbert in press) and input of nitrogen and phosphorus has increased about fourfold since European settlement (Moss et al. 1992; Neil and Yu 1996). While the inshore reefs are most impacted (Gabric and Bell 1993; Bell and Elmetri 1995; Brodie et al. 1997; Wachenfeld et al. 1998), the nutrients in river plumes may sometimes reach the outer reefs (Brodie 1996).


Bell and Elmetri (1995) strongly argue that the GBR lagoon is suffering from eutrophication, as indicated by increased phytoplankton density and a higher standing mass of macroalgae, though this assertion remains controversial (e.g., Walker 1991; Hughes and Connell 1999; Russ and McCook 1999)[14]. Nevertheless, algal overgrowth of corals is a recognised problem on many coral reefs (Szmant 2001). In some limited areas of the GBR region evidence of eutrophication is "indisputable" (Brodie 1997; Fabricius and De'ath 2001). However, Russ and McCook (1999), argue that the higher mass of algae on the inner reefs is not necessarily due to anthropogenic inputs of nutrients. They suggest that these areas may have fewer herbivores and that cyclones can greatly increase inshore production, probably through re-working of nutrients from sediments and increased river run-off. Brodie et al. (1997) monitored phytoplankton biomass associated with nutrient inputs in the Great Barrier Reef lagoon. They found chlorophyll levels were generally higher and more variable near the coast, but a compilation of 20+ years of data from the central GBR lagoon showed no evidence for a long-term increase. However, there were pronounced (2- to 4-fold) summer increases in chlorophyll at irregular intervals, emphasising the need for long-term monitoring studies to elucidate the natural patterns and extent of variation.


As a result of concerns about the effects of possible eutrophication of the GBR, the GBRMPA commenced an integrated research and monitoring program in 1991. One part of this was ENCORE (Enrichment of Nutrients on a Coral Reef Experiment), a large-scale in situ manipulative reef fertilisation experiment. In this study small patch reefs are being fertilized with nitrogen and phosphorus additions to assess the individual and combined impacts of nutrient enrichment on reef organisms (Steven and Larkum 1993). One aim was to separate the effects of elevated nutrient concentrations from other factors such as increased sedimentation, algal overgrowth and other pollutants (Koop et al. 2001). In summary, reef organisms and processes investigated in the ENCORE experiments were impacted by elevated nutrients, even at relatively low dosages. Dose level, or whether N and / or P were elevated determined particular impacts and were often species-specific. Moreover, these impacts were generally sublethal and subtle, and did not result in visible symptoms of a stressed coral reef (Koop et al. 2001).


Hughes and Connell (1999) argue that it is a mistake to attribute the current status of a coral reef only to present conditions (such as nutrient levels) when historical data do not exist because the current state could be due to a variety of causes such as past storm damage or long-term decline in grazers through overfishing. Experimental additions of nutrients resulted in algal blooms only when grazing fish were excluded (Hatcher and Larkum 1983) as grazing fish could ameliorate the increases in algal biomass following elevated nutrient levels. Thus while attention is often focused on a single stressor, the real world scenario involves a combination of complex, interacting stresses (e.g., overfishing plus altered nutrient loads, plus other natural stresses) (see also Section 6.9).


Eutrophication can have significant impacts on other marine ecosystems, including benthic communities. Heip (1995) discussed the possible effects of eutrophication on benthic dynamics, describing three successive states based on the amount of organic matter reaching the sediments: a) slight increases in biomass and few or no changes in species composition over the “normal” situation; b) large increases in biomass and replacement of “normal” species by opportunistic species; c) disappearance of benthic animal species and azoic sediments. Increases in benthic biomass and changes in species composition over decades attributed to eutrophication have been documented in a few cases (see Heip 1995). Mesocosm studies tend to show much more rapid changes (weeks to months), but often responses to experimental nutrient additions are inconsistent. Benthic fauna are important for benthic-pelagic coupling, bioturbation, benthic mineralization, and release of nutrients and dissolved organic matter in shallow waters. Consequently the disappearance of this fauna due to increased organic loading and/or anoxia could exert a significant influence on shallow water energy and matter cycles (Heip 1995). Significant changes in benthic communities as a result of eutrophication from aquaculture facilities have been documented in several cases (e.g., Everett et al. 1995; Lu and Wu 1998; see Section 6.11). 

Copyright © Environment Australia, 2002
Department of Environment and Heritage