The issues surrounding the potential effects of nutrients, including nitrogen, in relation to nature conservation in estuaries and coastal waters have been reviewed by Parr and Wheeler (1996), Scott et al (1999) and Parr et al (1999). The reader is referred to these documents for further details.

Nitrogen (N) can enter the marine environment from a variety of point and diffuse sources (see table below).

Nitrogen in coastal waters is derived predominantly from land run-off and direct discharges of sewage effluent from sewage treatment works (STW), although in nutrient-poor marine and estuarine areas, atmospheric deposition of NO_X and NH₄ may make a large contribution. Of terrestrial sources, run-off/leaching of nitrogen from agricultural sources (livestock waste and inorganic fertiliser) represents the major source in most catchments, but it is possible for sewage effluent to be the major nitrogen source in highly urbanised catchments.

As atmospheric deposition is the major nitrogen source in oceanic waters, it is sometimes assumed that it will represent a major source in tidal waters. However, this is not usually the case. The following nitrogen budgets (see table below) (Parr et al 1999) can be considered typical for many estuaries, with diffuse agricultural sources (livestock waste and inorganic fertiliser run-off) normally dominating the nitrogen inputs. Only in relatively small, highly urbanised catchments is the STW-derived nitrogen load likely to approach the diffuse source load. Nitrogen fixation by blue-green algae and bacteria in uncolonised sediment and saltmarsh typically accounts for a small proportion (typically <5%) of the N load to estuaries.

It is standard practice to monitor nitrate, nitrite and ammonium concentrations in tidal waters which, when added together, produce total inorganic nitrogen (TIN), an approximation of bioavailable nitrogen. However, some low molecular-weight organic nitrogen molecules, such as urea and some amino acids, are also taken up by algae and higher plants.

A wide range of concentrations are reported, with mean coastal water concentrations in England and Wales ranging from 0.07 to 1.85 mg l^-1 TIN (Parr et al 1999). Concentrations in the upper reaches of estuaries are similar to those found in river water, so concentrations here are often much higher, ranging from about 0.1 mg TIN l^-1 to 10 - 15 mg TIN l^-1. However, annual means hide a great deal of variability, with peak estuary concentrations occurring during autumn/winter and peak coastal water concentrations during winter. Thus, it is winter samples which are normally used for temporal and spatial comparisons of tidal water TIN concentrations.

MPMMG (1998) reported concentrations of ammonia, nitrite and nitrate for National Monitoring Programme sites in estuaries and coastal waters throughout the UK.

Annual instream nitrogen loads calculated using flow/monitoring data are presented for all Harmonised Monitoring Sites by Parr et al (1999). However, the N load measured at Harmonised monitoring sites may constitute less than half of the land-derived load (to smaller estuaries in particular).

The Environment Agency commissioned WRc to derive a classification scheme for nutrients in estuaries and coastal waters as part of their General Quality Assessment (GQA) scheme (Gunby et al 1995). The proposed GQA methodology for classifying TIN levels allows comparisons to be made using samples collected from different regions (salinities) of different estuaries, or different areas within the same estuary. This is because the GQA methodology assumes conservative behaviour for TIN and a standard concentration in marine waters, which allows the TIN concentration in the freshwater input to be calculated, provided salinity data are available. Estuaries can then be grouped according to the following class boundaries:

For a range of estuaries in England, this provides the following projected median TIN concentration in freshwater and GQA TIN class (see table below).

Different sources cite different nutrients as limiting in estuaries and coastal waters. Scott et al (1999) should be referred to for a more comprehensive review than is possible here.

As the freshwater input to estuaries from large lowland rivers usually has a nitrogen:phosphorus (N:P) ratio of >10, the water column is more likely to be P-limited than N-limited, particularly at the freshwater end of the estuary, but it appears that saltmarshes are usually N-limited. Although UK coastal waters have conventionally been described as nitrogen-limited, available data suggest that there are three major regions of coastline which are phosphorus-limited (i.e. the TIN:TRP ratio is >10; see Parr et al 1999). These regions extend from north of the Humber to the Essex estuaries, from the Solent to Dartmouth and around the Severn coastline from Padstow to Oxwich.

Nitrogen cycling in estuaries and coastal waters is a complex phenomenon and is outlined in Scott et al (1999) and in Parr et al (1999). Nitrogen is a major constituent of biota, so plants and animals (including plankton) are a sink for nitrogen. However, the largest sink for nitrogen remains the sediment, particularly in terms of organic nitrogen but, once in the sediment, this organic nitrogen can be broken down (mineralised) to produce bioavailable nitrogen, which is released back into the water column.

Nitrogen losses are achieved via denitrification (the conversion of nitrate to molecular nitrogen via ammonium). In estuarine and coastal shelf systems, this can amount to 7 - 54% of the total N or 11 - 57% of DIN (Dissolved Inorganic Nitrogen) (Nowicki et al 1997). The longer the residence time, the warmer the temperature and shallower the estuary, the greater N losses will be via denitrification.

The effects of non-toxic substances, such as nitrogen, on the marine environment can be sub-divided into direct effects (those organisms directly affected by changes in the concentrations of nitrogen species) and secondary effects (those arising in the ecosystem as a result of the changes in the organisms directly affected).

The terms nutrient enrichment and hyper-nutrification are used to describe the increasing concentrations of nutrients, including nitrogen, in the aquatic environment but do not relate to the consequences or effects of the increasing nutrient levels. The term eutrophication has been defined by the Environment Agency (1998) as Athe enrichment of waters by inorganic plant nutrients which results in the stimulation of an array of symptomatic changes. These include the increased production of algae and/or other aquatic plants, affecting the quality of the water and disturbing the balance of organisms present within it. Such changes may be undesirable and interfere with water uses.@ As such, it encompasses both the increasing nutrient levels and the resulting direct and indirect effects.

Using the monitoring techniques currently employed, it is very difficult to unequivocally make a case for nutrient enrichment having a deleterious effect on the quality of tidal water ecosystems, because of the many confounding factors (e.g. see Parr and Wheeler 1996). In particular, because high nutrient loads are often associated with high organic loads, it is often very difficult to distinguish between the effects of these two parameters. More ecologically relevant methods of monitoring the trophic status of estuaries are discussed by Scott et al (1999) and Parr et al (1999). The adoption of some of these novel techniques should make the identification and formal legal control of eutrophication a simpler process.

In the water column, the direct effects of increases in nitrogen compounds are the toxicity of ammonia and the response of algae and other aquatic plant communities using nitrogen compounds as nutrients.

Increasing nitrogen concentrations have been shown to be related to increasing phytoplankton standing crops (reflected in increasing chlorophyll-a concentrations). Thus, a number of authors have modelled chlorophyll-a and nitrogen levels using linear regressions (e.g. Gowen et al 1992). The extent to which phytoplankton standing crops increase in relation to increasing nitrogen concentrations is limited by factors, such as the availability of other nutrients (phosphorus and silicon, for diatoms), the level of turbidity and measures related to the time in which the increased nitrogen levels are available to the phytoplankton for growth. Consequently, relationships between nitrogen and chlorophyll-a concentrations are often weak, but may be improved by the use of correction factors to account for tidal flow/water velocity (e.g. Lack et al 1990). The use of similar factors to account for regional differences in turbidity around the UK coast (e.g. Parr et al 1998) could greatly increase the accuracy of such models.

Although it has been postulated that elevated nutrient levels are associated with increased occurrence of Phaeocystis (notably in the Adriatic, but also around the UK coast) and toxic dinoflagellate blooms, there is little evidence to support such a claim. The germination of spores lying on the sediment makes it much more likely that, following an initial bloom, blooms will reoccur in subsequent years.

Macroalgae found attached to hard surfaces in estuaries and coastal waters also respond to increasing concentrations of nitrogen in the water column, resulting in increased areal coverage and density. The most common macroalgal species involved are Enteromorpha spp. and Ulva spp.

Both nutrient and organic status influence the intertidal benthic diatom community, which may provide a good indicator of trophic status (e.g. Vos and de Wolf 1993, Parr et al 1999, Peletier 1996) - not in terms of biomass (chl-a), but rather in terms of species composition. However, the benthic diatom community consists of several sub-communities (those diatoms which live in interstitial water, those that are firmly attached to sediment granules, and those that live in frequently-desiccated areas of sediment). It is imperative that the same sub-community is monitored on all occasions. Epiphyte communities on macroalgae may similarly provide a sensitive indicator of trophic status.

The indirect effects of increased inputs of nitrogen are associated with changes in aquatic ecosystems resulting from the stimulation of algal and other plant communities in the water column and on the substratum.

While phytoplankton blooms are a natural occurrence, the nutrient enrichment of estuaries and coastal waters has been implicated in the more widespread occurrence and increased frequency and duration of such blooms. The plankton community includes consumers of phytoplankton, such as zooplankton which, in turn, are consumed by various metazoan animals (e.g. jellyfish and fish). The stimulation of phytoplankton will have knock-on effects for these consumers and change the community composition of the plankton with the possibility that biodiversity will be reduced.

Phytoplankton blooms can affect water quality during the growth and die-off phases. During the growth phase, the diurnal variation in dissolved oxygen can be exacerbated, such that, during the day, the water column can become supersaturated with oxygen as a result of the photosynthetic activity and, during the night, oxygen can become severely depleted due to respiration. Such fluctuations in dissolved oxygen can pose problems for invertebrates and fish and can lead to fish kills (see Section C5). During the die-off phase, the superabundance of phytoplankton cells in the water column and settling onto the substratum is a source of organic carbon for aerobic bacteria which rapidly decompose this material but, in doing so, can exert an oxygen demand on the water column and on the sediments resulting in severe oxygen depletion. This can lead to sublethal and lethal effects on invertebrates and fish. These changes in water quality are likely to be greatest in semi-enclosed bodies of water with long retention times and where stratification of the water column occurs.

Phytoplankton blooms can contribute to an increase in turbidity in the water column which reduces the light availability to macroalgae and plants growing in the photic zone and resulting in a reduction in the depths of colonisation for several species (Parr et al 1998, Birkett et al 1998).

Repeated phytoplankton blooms can lead to severe degradation of the marine environment with potential adverse consequences for birds and sea mammals as the diversity and abundance of food organisms change.

The stimulation in the growth of macroalgae, particularly Enteromorpha and Ulva, in intertidal areas can result in the formation of an extensive cover of algal material (an algal mat) on the surface of exposed sediments. Above a critical standing crop, the increased amount of organic material and the reduced exchange of water between the sediment and the water column result in deoxygenation of the sediments. This can lead to a change in the infaunal benthic community and, in severe cases, the death of many benthic invertebrate species. Extensive coverage of intertidal sediments can severely diminish the feeding areas for fish and birds.

For a full range of possible direct and indirect effects of nutrient enrichment, the reader is referred to Scott et al 1999.

Potential effects include:

• toxicity of ammonia to invertebrates and fish in the water column;
• stimulation of phytoplankton growth in the water column of estuaries and coastal waters;
• stimulation of macroalgae, particularly Enteromorpha and Ulva spp., on the substrata of estuaries and coastal waters;
• perturbation of the plankton community, including zooplankton, other invertebrates and fish, as a result of repeated phytoplankton blooms with the potential to reduce biodiversity;
• increased fluctuation is dissolved oxygen concentrations in the water column during the growth phase of a bloom with the potential for sub-lethal and lethal effects on invertebrates and fish;
• potential for depletion of oxygen concentrations in the water column and sediments as a result of the die-off of phytoplankton blooms with the potential for sub-lethal and lethal effects on invertebrates and fish;
• contribution to increased turbidity in the water column and reduction in light availability to macroalgae and other aquatic plants growing in the photic zone;
• reduction in oxygen availability in intertidal sediments under algal mats with the potential for sub-lethal and lethal effects on infaunal invertebrates and reduced feeding areas for fish and birds;
• potential for severe degradation of the ecosystem with adverse consequences for sea birds and Annex II sea mammals.

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