1.0 Introduction
1.1 Eutrophication
The trophic state or fertility of water bodies is defined by factors relating to autotrophic production such as algal biomass, water column nutrients and water transparency (Dodds 2007). There are three classes of trophic state, defined by Dodds (2007) as; oligotrophic (unproductive), mesotrophic (intermediate productivity) and eutrophic (highly productive). A waterbody with an increase in the factors relating to autotrophic production will be classed as eutrophic due to excessive nutrient loading. Mason (2002) has described eutrophication as the “enrichment of waters by inorganic plant nutrients”, however, definitions do vary as do the processes that cause eutrophication.
The concept of eutrophication has appeared in scientific papers since 1918 with an increasing understanding of its effects in the freshwater ecosystem (Phillips et al 1999). The excessive nutrient loading leads to high levels of biological activities which therefore require utilisation of nutrients. Primary producers such as algae will use more and more of these and biomass will develop (Hendry et al 2006). The increase of nitrogen and phosphorus increase the abundance of aquatic plants and primary productivity but it is the timing of the addition of excess nutrients that determines community diversity (Dobson et al 2009). This in turn affects the aquatic ecosystem and the food-web when they decompose by increasing biological oxygen demand leading to a decrease in available oxygen and eventually the death of fish and animal life (Hendry et al 2006). Cyanobacterial blooms are another consequence of eutrophication which can have a devastating effect on the aquatic ecostucture. This will be discussed in further detail throughout the introduction.
The past two centuries have experienced excessive anthropegenic inputs which in turn have led to undesirable changes in the environment with almost a half of the lands surface being transformed (Merrington et al 2002). The term cultural eutrophication refers to the increase in nutrients due to these anthropogenic inputs whereas natural eutrophication occurs via non-human processes (Mason 2002).
Cultural eutrophication is a problem because it is a form of pollution; Holdgate (1979, cited in Mason 2002) has defined pollution as;
“ The introduction by man into the environment of substances or energy liable to cause hazards to human health, harm to living resources and ecological systems, damage to structure or amenity, or interference with legitimate uses of the environment”.
The sources, effects and problems of cultural eutrophication will now be discussed further.
1.1.1 Sources of nutrients
Nutrients in the form of phosphates and nitrates are essential for plant growth and biological activity occuring naturally at harmless levels in the environment. Phosphorus is essential for DNA formation, energy storage and transfer in all organisms while nitrate is essential for growth (Merrington et al 2002). These combined in excess are ideal for excessive algal growth due to nutrient supply being the limiting factor, especially phosphorus, and therefore the controlling factor effecting primary production in freshwater ecosystems (Van der Molen et al 1998).
However, with reference to cultural eutrophication which leads to excess nutrient loading it is important to identify the sources of these nutrients. As well as the consequences for the ecosystem, there are statutory requirements that need to be adhered to with reference to water quality, nitrate limits in water, duty of care to the environment and commercial and amenity values. Point sources such as phosphorus from sewage treatment works (STW) (Jarvie et al 2008) and nutrients due to poor storage of slurry or silage (Hooda et al 2000) are easier to identify and control than non-point sources such as fertilisers, due to changes in farming practices and weather patterns (Withers et al 2002). The decline in traditional mixed farming methods with an increase in specialist arable and crop farms is another harmful effect of agricultural intensification. Intense arable cropping involve periods of incomplete crop cover which encourages leaching events due to excess nutrients unable to be absorped by crops (Merrington et al 2002).
The majority of nutrients from urban sources are derived from STW, storm drains and industrial waste, while nutrients from rural areas include agriculture, rural dwellings and forest management (Mason 2002). Jarvie et al (2010, 2008) found significant point sources included rural areas who used septic tank systems without sewage treatment works. There have been numerous studies of watercourses in order to identify timelines to determine when eutrophication has begun. The majority have focused on analysing sediment samples to determine when the increase in nutrient levels started and then associated to historical events such as changes in the use of landscape or the opening of sewage treatment works. Sediment layers can indicate seasonal inputs and historical inputs and are a reliable and direct proxy of changes in aquatic productivity during eutrophication (Yuehan et al 2010).
The UK has had a major increase in the number of livestock increasing from 108,000,000 in1940 to 188,000,000 in 1987 (Hooda et al 2000). A case study of Slapton Ley, Devon involved a reconstruction of the diatom analysis of sediment cores which suggested that Total Phosphorus concentrations began to increase at c. 1910 which indicated the change from arable farming to more intensive livestock production and agricultural intensification (Mainstone et al, 2008). Sediments can also determine the biogenic silica (BSi) that has been absorbed by diatoms, up to five percent of the biogenic silica produced annually by diatoms can be buried in sediments ( Schelske 1985). Yuahen et al (2010) found that sediment core analysis showed a rapid increase in BSi content due to the rapid urbanisation in the 1950s until 1970 when there were changes in external nutrient loadings and trophic status. High Nitrogen:Phosphate (N:P) ratios were found between 1930 and 1950 indicating the beginning of eutrophication which suggested an efficient phytoplankton uptake of nitrogen when the Phosphorus supplies are also abundant. Between 1950 and 1970 the N:P ratios were low which indicated nitrogen was being depleted along with high levels of P inputs. Peaks in sediment cores have indicated that high levels of phosphates have been followed by low levels of BSi after Si was minimal (Yuahan et al 2010). Yuehan et al (2010) found a peak of BSi in the 1970s from spring diatom blooms due to higher levels of phosphates from sewage and detergents following urbanisation.
1.1.2 Water Pollution from agriculture and run-off
Since the 1930s the agricultural industry has become increasingly intense with the availability of pesticides, herbicides and fertilisers enabling high nutrient levels to be maintained in agricultural soils the whole year. This use of landscape has many effects on the environment such as loss of habitats as well as deterioration in soil, water and air quality (Merrington et al 2002).
The Agricultural Act was introduced in 1947 to instigate and regulate a market and price structure for farmers produce. This increased agricultural production, with the advances of genetic modification and precision farming intensifying this market (Merrington et al 2002). The Common Agricultural Policy was an EU initiative to support agricultural practices following objectives from the 1957 Treaty of Rome (Merrington et al 2002, Mainstone et al 2008). This was a time when food was in short supply and therefore created an economic climate concerning food production with larger, intense farms specialising in crop and livestock production (Merrington et al 2002).
1.1.3 Nitrates and nitrogen loss
Depending on the different farming methods and practices, different ways of leaching can occur. The efficiency of nutrients is never 100% and grassland systems have been known to only use 10% of N efficiently (Merrington et al 2002).
The effect of diffuse pollution on the quality of surface water, and therefore surface run-off has resulted in excess nutrients primarily in the form of organic manures arising from livestock production. It has been estimated (Isermann 1990, cited in Smith et al 2001) that between 37 and 82% of nitrogen and 27 and 38% of phosphorus into surface waters comes from agriculture. Kronvang et al (1996) found that of 270 Danish streams, 94% of the nitrogen loading and 52% of the phosphorus loading primarily arose from diffuse sources of agricultural practices.
Nitrogen and Phosphorus losses vary between different farming practices and landscape characteristics such as soil type and topography are significant factors when it comes to nutrient loss (Withers et al 2002). Due to the solubility of nitrate, any NO3- made available in soil is vulnerable to losses. This occurs when plants stop growing as nitrogen uptake ceases and therefore rain or irrigation can cause leakage of nitrogen into drainage water and surface run-off (Withers et al 2002). Although nitrogen occurs naturally in soils it is the transformations within the soil (soil N cycle) that influence the amount lost from the soil (Merrington et al 2002).
Fertilisers in the form of animal manure are spread in autumn and winter to enable the process of mineralisation to take place to leave behind large residues of nutrients ready for uptake from crops (Merrington et al 2002). 125 kg ha -1 of nitrogen, 14 kg ha -1 of phosphorus and 35 kg ha -1 of potassium are received by fertilisers on arable and managed grassland areas in the UK (MAFF et al 2000).
The timing of the application of manure is a significant factor when considering nutrient loss. Studies carried out by Smith and Chambers (1998, cited in Withers et al 2002) showed that there is a high risk of leaching when high soluble nitrogen content manures were spread in September to November, whereas the best times to spread manure are spring and mid-late winter if the soils are retentive or in a drier area, soil nitrate concentrations may exceed the 50 mg l-1 if the application of fertiliser has been poorly timed (Withers et al 2002). The slurry and manure can contain 70% of nutrients ingested by cattle and 80% of nutrients ingested by sheep which is why there are excess nutrients in run-off when rainfall occurs immediately after manure applications to soil (Withers et al 2002).
1.1.4 Farming practices
Elevated levels of NO3- in groundwaters have been linked to the initial ploughing of grassland. This was particularly apparent in the 1940s -1950s when grassland was being cultivated to make way for arable cropping. When soils are cultivated an increase in NO3- is made available due to stimulation of the mineralisation stage in the N cycle (Merrington et al 2002).
Livestock farms generally produce excessive amounts of nutrients in the form of manure. Some crops such as bread-making wheat require a large supply of nitrogen late in the growing season whereas too much nitrogen is detrimental to others, horticultural crops tend to leach from the organic mineralisation stage as do arable soils and winter cereal crops (Merrington et al 2002)
1.1.5 Phosphorus
Phosphorus is a major contributor to eutrophic waters and algal blooms with Morse et al (1993) estimating that 43% of phosphorus inputs into surface waters in the UK come from agriculture. Phosphorus is generally adsorped to soil particles and therefore lost, particularly in autumn and winter, when attached to eroded soil particles in surface run-off specifically in arable areas and heavily trampled grassland due to a lack of plant cover (Withers et al 2002, Mainstone et al 2002). When events such as rainfall occur, phosphorus losses occur, especially when the topsoil is rich in phosphorus for example from receiving large amounts of manures for prolonged periods and where fertilisers and manures are applied together (Withers et al 2002)
It would seem point sources such as STW are becoming increasingly controlled in order to control phosphorus levels and keeping them down to background levels, however diffuse sources do also need to brought under control due to the strong positive correlation between total phosphorus concentration and algal biomass and continual input throughout the year when ecological sensitivity is at its highest (Dobson et al 2009, Jarvie et al 2008, Mainstone et al 2002 ).It must be remembered that point sources enter water courses all year round and with the high bioavailability from STW effluents they can have an immediate impact. Human waste, detergents and trade wastes enter the sewerage system which then receive a secondary treatment that becomes bioavailable as orthophosphate (Mainstone et al 2002).
1.1.6 Internal loading of nutrients in sediments
Internal loading of phosphorus may occur in river bed sediments which can continue to be taken up by benthic algae and the roots of higher plants (Mainstone et al 2002). As an example, Lake Trummen in Sweden experienced thirty years of sewage effluent loading. This resulted in 8mm year-1 of black sulphurous mud which stored phosphorus giving an internal load of 177kg phosphorus year-1 as opposed to the reduced external load of 3kg yr-1 from groundwater, fluvial and atmospheric inputs (Mason 2010, Smith et al 1999). The main problem with silty substrates is that the phosphorus release rate is enhanced due to anoxic conditions (Mainstone et al 2002)
1.2. The Impacts and Effects of Eutrophication
1.2.1 Economical effects
As can be seen in Table 1 the estimated costs per year of eutrophication are spread across a number of different areas having to deal with the effects of eutrophication. Concentrations of nitrate above 50mg L -1 have been linked to human health problems such as ‘blue-baby’ syndrome (Hooda et al 2000)
Table 1: Costs of eutrophication in the USA and England and Wales (table taken from Moss 2010)
1.2.2 Ecological effects
The effects of eutrophication on algal communities has concluded that primary production is increased, with changes in species composition of suspended algae and periphyton (Smith et al 1999). The presence of nutrients at excessive concentrations can severely disrupt the aquatic ecosystem leading to increased algal biomass which reduces the amount of light into the water column, algal blooms, a decrease in aquatic diversity geared towards plants that require higher nutrient levels and the death of organisms living in the affected water body (Mainstone et al 2002). The increase of nitrogen and phosphorus increase the abundance of algae which eventually outcompetes the growth of submerged macrophytes (Moss 2010). This in turn affects the aquatic ecosystem and the food-web. The lack of submerged macrophytes which usually store nutrients in the plant biomass, will lead to higher levels of nutrient availability to phytoplankton. Plant-inhabiting zooplankton and invertebrates do not have suitable refuges which in turn will lead to an increase in predation and a decrease in their primary predation indicative of an increase in algal density and therefore algal blooms (Moss 2010). In shallow lakes the water may become turbid due to the increase in phytoplankton density depleting oxygen. Organisms tolerant of lower oxygen concentrations will dominate leading to the exclusion of certain fish groups such as white fish and salmonids (Moss 2010). The loss of submerged macrophytes can severely affect the spawning and living habitat for fish.
Algal blooms
Cyanobacterial algal blooms consist of blue-green algae and are fairly widespread in eutrophic lakes due to the increase in anthropogenic activities, such as intensive agricultural practices and changes in land use of the surrounding areas (Figueiredo et al 2006). The toxic consequences of the different strains of these blooms are associated with negative economical, ecological and public health implications.
The algal blooms are made up of nitrogen fixing algae, although phosphorus has been reported as the nutrient of greatest concern when it comes to freshwaters, nitrate levels are arguably, of equal importance (Mainstone et al 2008). Heisler et al (2008) summarised a discussion held by the Environmental Protection Agency in 2003 regarding the presently understood relationships between eutrophication and harmful algal blooms that were either toxic or high biomass producers causing hypoxia and anoxia. The report indicated a strong relationship between phosphorus inputs and harmful cyanobacterial blooms in freshwaters.
Cyanobacteria have a physiological adaption that enables them to succeed in nitrogen depleted conditions depending on other environmental conditions (Hooda et al 2000) Their ability to store phosphorus and regulate buoyancy through gas vesicles can deplete the abundance of diatoms and chlorophytes while being extremely toxic to cladocerans (de Figueiredo et al 2006).
The toxins produced by cyanobacteria will also depend on nutrient form and whether the nutrient forms are organic or inorganic (Heisler et al 2008). There are over fifty known cyanobacterial blooms with each having different strains of toxicity. The phytoplankton community dynamics can indicate the toxicity of the bloom due to the chemical substances they may produce (Figueiredo et al 2006). Douterelo et al (2004) cultured different species of benthic cyanobacteria in order to monitor water quality in running waters. This confirmed that the species from the order Oscillatoriales were most abundant in eutrophic waters with many of these species being used as biological indicators of organically polluted waters. The Oscillatoriales are a toxic cyanobacterial species as well as Microcystis (Figueriedo et al 2006).
Microcystin is a hepatotoxin which can kill fish, birds, wild animals, livestock and humans (Choi et al 2005). The ecological role of Microcystins is still being studied but the main hypothesis for it producing toxins is to act as a deterrent against grazers (LeBlanc et al 2004). Organisms are known to directly or indirectly affect other organisms through releasing certain compounds. This alleopathic effect was investigated by LeBlanc et al (2004) concerning the ability of Microcystic Aeruginosa to outcompete the macrophyte Lemna Gibba. The experiment concluded that the toxic cyanobacterium was not significantly toxic to Lemna Gibba and despite an expectation of an increase in toxic production it actually decreased by day seven. There is evidence that Microcystic Aeruginosa can effectively inhibit zooplankton and algal species. Jang et al (2003) discovered that the toxic cyanobacteria produced significantly more toxins in the presence of zooplankton. This was believed to be an allelochemical response due to chemical signals from the zooplankton stimulating greater amounts of toxins from Microcystic Aeruginosa in the presence of a predator.
Microcystis aeruginosa was reported to have over forty strains of toxicity in Vela Lake, Portugal which inhibited the development of phytoplankton and zooplankton. The levels of diatoms and green algae decreased during the Microcystis aeruginosa bloom while dissolved oxygen levels depleted and there was an increase in ammonium levels. Due to the phosphorus storage abilities of Microcystis aeruginosa, a depletion in phosphates can still maintain a high density for a few months following a bloom provided there are continuous sources of nitrogen (Figueiredo et al 2006). Jöhnk et al (2008) noted that Microcystic blooms increased in population during hot summers as opposed to colder summers due to requiring a higher temperature optimum than diatoms and green algae. This increase in population results in dense surface blooms providing competition for light against other phytoplankton.
1.3 Nutrient Control in Lakes and Catchment Areas
As discussed previously it is the limiting nutrient that causes problems in eutrophic water bodies and therefore a reduction of this nutrient should reduce the algal density. Phosphorus supplies are easier to identify and control than nitrogen as they generally only come from a few point sources such as STW. However if the limiting nutrient is nitrogen and has increased in areas where phosphorus has accumulated naturally it would be necessary to control nitrogen (Moss 2010).
Legislation and changes in agricultural practices have been put in place but it is the present intensive farming that will need to change in order for control of nitrogen to become effective (Moss 2010). To restrict nutrient loss in the short to medium term it has been suggested to reduce livestock numbers, convert arable land to low-input grassland , change the timing of applying manures, avoid high risk crops in vulnerable areas and use over-winter cover cropping which can reduce nutrient leaching by 50% (Withers et al 2002).
The Environment Agency’s national strategy for eutrophication control (2000, cited in Mainstone et al 2002) has an aim of having phosphorus levels between 0.02 and 0.2mg L-1 (mean Soluble Reactive Phosphorus) in different types of rivers such as heavily enriched rivers or lowland rivers on clay and alluvium. Mainstone et al (2002) suggest that there are four main areas to controlling agricultural pollution, these being “nutrient management, soil management, run-off management and crop and livestock management”.
Phosphate stripping is another form of controlling phosphorus loading from sewage effluents. This approach is effective but if the affected lake has received phosphorus loading from agricultural practices it can take 5-20 years for this approach to significantly reduce this form of nutrient loading due to internal phosphorus loading in sediments (Mason 2010). It is therefore important to identify the internal and external loads of phosphates released (Mason 2010). Surface layers of sediments have been dredged to reduce the size of this internal reservoir as well as gravels being artificially cleaned if silted up with phosphorus-rich silt, these methods do depend on local circumstances due to the impact dredging can have on some instream habitats (Mainstone et al 2002). Attempts have been made to remove or seal sediments when external sources of phosphorus cannot be controlled but sediment behaviour is still not entirely understood (Mason 2010).
Phosphorus in run-off can also be avoided by the use of buffer zones, changes in the direction of cultivation on steep slopes and modifying field size (Withers et al 2002).
Buffer zones are used to remove nitrogen from run-off and leaching that may otherwise enter a water body. Other forms of control have included new approaches towards landscape management and environmental conservation and through legislation.
Biomanipulation has been used as a form of control in water bodies. Top-down biomanipulation “reduces the predation of the organism which has the largest impact on the organism causing the problem” while bottom-up manipulation usually occurs when the nutrient levels have already been reduced (Holley 2002). The water-level of water bodies has also been manipulated (Wilson et al 1998).
1.4 Legislation
The Water Framework Directive requires that all sites should achieve ‘good status’ by 2015. River Basin Management Plans were published in 2008/2009 in order to manage problems across point and diffuse sources of pollution in water bodies from catchment areas (Mainstone et al 2008). In addition groundwater needs to meet a good chemical and quantitative status through the designation of protected Areas (PAs). Many agricultural streams should have a mean annual concentration below 100µg P L -1 due to the Water Framework Directive requiring improved riverine ecology (Jarvie et al 2008) while UK studies associated with the Habitats Directive have set limits of > 30µg L-1 for annual mean of soluble reactive phosphorus (Kelly et al 2008) The UK Code of Good Agricultural Practice for the protection of water limits the application of slurry to 50 m3/ha (MAFF 1998, cited in Smith et al 2001).
Excess nitrates are a cause for concern due to EC drinking water limits which should not be above 50mg L-1. To protect the quality of drinking water nitrate levels need to be below 50mg/L (FWAG 2009) as the water will be unsuitable for drinking without further treatment (Withers et al 2002). Nitrate Vulnerable Zones (NVZs) now cover up to 70% of England with an emphasis on restricting inputs and spreading manures at the right time of year to avoid a risk of oversupply of nutrients in the current year (Withers et al 2002).
Since January 2009 new rules came into force with the Nitrate Pollution Prevention Regulations 2008 providing the legislation for owners of land in a NVZ . These Regulations are implemented under the auspices of the European Community’s Nitrates Directive which designate these zones. The Regulations are enforced through the Environment Agency.
The integrated Pollution Prevention and the Control Directive is aimed at targeting large livestock practices while the EU Habitats Directive targets areas with special conservation status in order to maintain or restore favourable environmental conditions for conservation (Withers et al 2002, Mainstone et al 2002). The Urban Waste Water Treatment Directive concerns identifying sensitive areas (waters that are or are in danger of becoming eutrophic) in order to provide the appropriate treatment facilities of phosphorus and nitrogen removal when the sewage treatment works serve more than a population of 10,000 (Mainstone et al 2002). From 2002 -2005 seventy-four sewage treatment works that affected SSSIs in the UK had phosphorus removal systems fitted and the Government Public Sector Agreement (PSA) have a target of bringing 95% of SSSIs into favourable, or unfavourable but recovering condition by 2010 (Mainstone et al 2008). However north-western Europe have very strict controls on phosphorus removal with Denmark, the Netherlands, Finland and Sweden designating their whole area as sensitive (Mainstone et al 2002).
1.5 Conclusion
The interaction of nutrient enrichment and anthropogenic impacts within water bodies has many ecological impacts leading to many obligations and pressures to have a duty of care to the environment. The precautionary approach to maintain and restore habitats and species has acted as a template in order to implement controls and conservation strategies to enhance the ecological status of freshwaters and the environment.
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