Eutrophication is the build up of inorganic plant nutrients in the water body. The effects, in extreme circumstances, can result in reduced water clarity, lowered dissolved oxygen levels, and toxic water quality. The causes, effects and monitoring of eutrophication are considered in turn.

The nutrients of primary concern are nitrates and phosphates, and these enter the seawater by a variety of routes: outflow in rivers, direct discharges of sewage and industrial effluents, and atmospheric input all contribute. The concentrations of these nutrients have increased substantially in many British coastal areas in recent years, and are a matter of increasing concern. Thus in the Irish Sea nutrient levels have roughly doubled over the past forty years (Allen et al., in press), and some of the symptoms of eutrophication described below are becoming increasingly evident (Shammon et al., 1997).

The primary effect of eutrophication is to stimulate algal growth, both benthic macroalgae and the microscopic phytoplankton. The adverse effects of excess macroalgal growth are largely aesthetic, caused when increased amounts are cast up on the strandline, or when dense algal growth carpets intertidal areas. However, the effects of phytoplankton proliferation are more serious. Phytoplankton blooms fall into two categories. ‘Nuisance’ algae (e.g. Noctiluca, Phaeocystis) can create problems by discolouring the water, creating aesthetic nuisance, and more severely by de-oxygenating the water and killing fish and benthic organisms. ‘Toxic’ algae such as Dinophysis can be taken up by shellfish which if eaten may produce diarrhoetic shellfish poisoning (DSP), whilst Pseudonitzchia can induce amnesic shellfish poisoning (ASP). Both of the above genera were recorded in the Irish Sea in 1997 (Shammon et al., 1997), and there are established guideline levels and monitoring procedures (Anderson, 1996).

Although the process of eutrophication is unambiguously linked to increased nutrient levels, there is no clearcut correlation between specific nutrient concentrations and degree of eutrophic phenomena. Nutrient levels associated with serious eutrophic damage in one area may be without obvious effect in others. The problem is complicated by the wide fluctuations in surface nutrient levels during the year, from winter maxima to virtual absence in the summer ( see Kennington et al., 1997 for a recent case study in the Irish Sea). For management purposes the only appropriate strategy is to take account of the standards proposed by the Comprehensive Studies Task Team (CSTT) of the U.K. Government in relation to meeting the requirements of the E.U. Urban Waste Water Treatment Directive (UWWTD). An area is considered to be ‘hypernutrified’ if the winter nutrient concentrations exceed 12 mmol DAIN (dissolved available inorganic nitrogen) m^-3 in the presence of at least 0.2 mmol DAIP (dissolved available inorganic phosphorus) m^-3 (CSST, 1997).

Since by definition CFT communities are essentially animal dominated, the effects of eutrophication will be indirect. One effect of eutrophication will be the way it influences the growth of benthic macroalgae, which may influence the level of the boundary between the infralittoral and the circalittoral. Improved macroalgal growth might be expected to lower this boundary, but at the same time increased phytoplankton density will reduce light penetration, perhaps more than compensating for any improved growth. Observations confirm that eutrophication does in fact raise the lower limit of macroalgal growth (Kautsky et al., 1986; Michanek, 1972; Svane & Gröndal, 1988) - in the Baltic from 11.5 m in 1944 to 8.5 m in 1984 (Kautsky et al., 1986). On the Swedish west coast subtidal rocky areas previously algal covered had been taken over by mussels in 1988 (Lundälv, 1990). Large algae are also affected by the improved competitive advantage of ephemeral filamentous algae in higher nutrient concentrations (Lundälv et al., 1986; Rueness, 1973; Wallentius, 1984). It is unlikely that effects on the macroalgae will have major implications for the CFT biotopes.

Changes in the phytoplankton are more likely to produce impacts. Increased phytoplankton densities will change the food supply for the predominantly filter feeding CFT species - the effects will be uncertain. Blooms of toxic algae may affect survival of CFT species, perhaps particularly in their planktonic larval stages. Algal blooms are often considered a near-surface phenomenon, and more likely to pose a threat in sheltered conditions. However, major effects of toxic algal blooms (especially of the species Chrysochromulina polylepis and Gyrodinium aureolum) have been reported from exposed sites along the Norwegian and Swedish Skagerrak coasts down to depths of 25-30 m (Bokn et al., 1990; Lundälv, 1990, 1996). So neither depth nor exposure necessarily offer protection. As well as the toxic impact of blooms, deoxygenation of the water will clearly have adverse effects. CFT communities in exposed high energy situations (which includes the majority in British waters) are probably at little risk from this, but those in semi-enclosed locations may be (Marchetti, 1992), and the risk of bottom water deoxygenation should be considered. Low bottom oxygen levels in Scandinavia have been linked to eutrophication (Lundälv, 1990).

Overall eutrophication poses a variety of threats to CFT communities, though it is currently impossible to relate risk directly to nutrient levels, and probably SAC management can exert little control over nutrient inputs. As for other forms of pollution, shallow water and sheltered area CFTs, such as those in sea lochs and rias, will be most at risk. A monitoring programme should include an assessment of eutrophication related phenomena.

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