Water quality issues in the British uplands

Applied Geography (1991), II, 201-214

Water quality issues in the British uplands

Gordon N. Mitchell

School of Geography, University of Leeds, Leeds LS2 9JT, UK

Abstract

A review of the major issues in UK upland catchment water quality research is presented. For each issue, an attempt is made to identify why it is perceived as a problem and to examine its extent and historical trends. Causal mechanisms are discussed and possible control methods investigated. Issues examined include acidification, mobilization of toxic metals, discoloration, nutrient loss and sedimentation.

Introduction

The British uplands, areas over 250 m above sea level, occupy 7.3 x lo6 ha, approximately one-third of the total UK land area. They are concentrated in Scotland, Wales, northern England and the west country, and are characterized by high rainfall, low temperatures and acid soils. These conditions limit the growing season and agricultural productivity. Land use is mainly restricted to rough pasture supporting stocking densities up to 2.5 ewes ha- I, improved pasture supporting 6 ewes ha- I and coniferous plantation. The EC less favoured area directive (75/268/EEC) has sus- tained the level of improved pasture, while government policy has greatly increased the post-war rate of conifer afforestation through a system of planting subsidy and tax incentive. The extent to which these systems of land use subsidy are ecologically or financially sustainable has not been defined (Boon and Kay 1990).

Due to their topography, high rainfall, remote location and low-intensity land use, the British uplands are well suited to the gathering and storage of water. Upland reservoirs supply approximately one-third of water consumed in the UK and fulfil a vital function in river regulation. Upland water resources must therefore be considered as a nationally important asset.

The main thrust of UK upland catchment research has until relatively recently centred on questions of quantity. Pioneer studies in Lancashire suggested that afforestation could reduce the yield of a catchment by as much as 42 per cent (Law 1956). Further research, including a ten-year paired catchment study at Plynlimon, Wales, demonstrated that Law’s observations were correct and that catchment afforestation could substantially reduce water yield (see, for example, Rutter 1963; Calder 1976, 1979; Newson 1979). These studies brought water resource agencies into direct conflict with forestry interests, which were dedicated to the establishment of a national timber reserve, over the strategic use of the uplands. This conflict remains to be resolved satisfactorily and in the meantime the emphasis of UK upland catchment research has shifted from questions of quantity to those of quality.

Land use and climate affect the quality of upland water, changes producing a reduction in quality that adversely affects potable supplies, freshwater ecosystems and the recreational value of surface waters. Potable supplies are particularly susceptible to changes in upland use due to the traditional method of quality protection. This consisted of catchment protection, where land use and access were

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202 Water quality issues in the British uplands

strictly controlled, coupled with long storage which necessitated only simple filtration and chlorination near the point of use. However, increasing consumption has placed greater demands on upland water resources, thus reducing storage periods, while recent privatization of the industry has produced pressure to change to land uses that are not fully compatible with water gathering. The relatively low cost of upland sources has promoted their use in preference to groundwater and river abstraction wherever the development of the supply network permits a choice of source. These pressures reduce the ability of present water treatment works to cope with deteriora- tions in water quality, which can directly degrade aquatic and riparian ecosystems both above, below and within impoundments. Finally, water quality reductions can reduce the amenity value of surface waters for angling and immersion sports.

In this paper a review of the major issues in UK upland catchment water quality research is presented. For each issue, an attempt is made to identify why it is perceived as a problem, and to examine its extent and historical trends. Causal mechanisms are discussed and possible control methods investigated. The review concludes by suggesting ways in which upland catchment research and water resource management may proceed in the future.

Acidification of surface waters

Acidification of surface waters in parts of Europe, North America and Canada has been cause for growing concern in recent years. This concern arises from the possible adverse impacts acidification has on the ecological and economic value of lakes and rivers. Acid waters may also be significant in terms of potable water and therefore public health.

The biological impacts on fish are perhaps the most extensively researched due to their clear economic, conservation and recreation value. Due to the variable reliability of information on changes in fish populations and the poor understanding of fish ecology in waters prone to acidification, it is difficult to quantify accurately the impact of acidification on those populations. However, acidification is thought to have been responsible for a decline in salmon (Salvo salar) in Wales (Warren 1989) and brown trout (Salvo trutta) in parts of Scotland and Wales (Maitland et al. 1986; Milner and Varallo 1990). Acidification is thought to be responsible for the total loss of salmonids in some Swedish and Norwegian lakes (Almer et al. 1974; Sevaldrup et al. 1980), while short-lived acidic flushes have been responsible for catastrophic fish kills in the UK (Robinson 1986).

Evidence suggests that acidification affects all trophic levels to some extent. Microbial and fungal activity may be reduced in acid waters, resulting in an accumula- tion of undegraded organic matter (Traan 1980; Chamier 1987), affecting overall productivity. This effect is likely to be most significant in lakes where disturbance by current and feeding by invertebrates are reduced. Primary production in upland lakes and streams is thought to be adversely affected by acidification, through indirect processes. Acidification is only thought important in oligotrophication when it reduces recycling of nutrients from sediment (Hendrey et al. 1976) or removes phosphate by precipitation with aluminium (Dickson 1978). Macroinvertebrate populations are less diverse in acid waters (see, for example, Wright et al. 1984) and as a food source may affect the distribution of higher trophic levels, such as birds and fish. The reproductive success of some amphibians, such as the common frog and natterjack toad, may be affected by acidification due to the availability of food or lethal and sub-lethal toxic effects of acidity, particularly on embryos (Pough 1976; Dunson and Connel 1982). Semi-aquatic birds and mammals can also be affected by

Gordon N. MilchelI 203

acidification. The reduction in fish populations may affect the distribution of piscivores with high conservation value, such as osprey and otter (Ormerod and Tyler 1987; Mason and MacDonald 1987). Low pH waters present no direct risk to people, but acid waters have a propensity for dissolving toxic metals from soil and pipework which, if not removed from potable supplies, may have adverse effects on public health.

Surface water has been divided into three broad categories (Warren 1986): waters which are permanently acid, waters which are occasionally acid, and those that are never acid. Acid is defined as < pH 5.6. The acidity of water is strongly regulated by alkalinity. Permanently acid waters have alkalinity close to zero, while waters which are never acid have high alkalinities (> 5 mg 1-r CaCos) and are usually above pH 6.0. Waters with low alkalinity (< 5 mg 1-r CaCos) are poorly buffered and are therefore most susceptible to acidification.

The distribution of known acid waters, and of waters thought to be susceptible to acidification in the UK, has been determined by the UK Acid Waters Review Group (Warren 1986). These distributions are determined by geology (Kinniburgh and Edmunds 1984), soil type (Catt 1985) and topography. Waters most susceptible to acidification are associated with non-calcareous lithology and soils of low buffering capacity. Vulnerable areas include much of Wales, Scotland, and Northern Ireland, while in England they are confined to the north and west and localized southern areas.

Evidence for acidification cannot reliably be found in archives held by regional river authorities, water utilities and river purification boards. These records are typically too short (< 20 years), only include pH (which is often poorly measured in low ionic strength waters) and are usually not determined for the upper reaches of rivers where acidification is most evident. However, strong evidence of acidification has come from changes in diatom assemblages in lake sediments (see, for example, Batterbee et al. 1985; Warren 1989). Diatom distribution in freshwater is strongly regulated by pH and factors that co-vary with pH such as alkalinity and aluminium. These analyses indicate initial acidification of lakes due to the development of organic soils following the last ice age. However, during this period no waters fell below pH 5.6. A definite trend towards lower pH begins around 1850, with pH values falling by 0.5-l -5 units from their pre-1850 levels to 4-5-5 ‘5.

The timing of the onset of acidification varies spatially in relation to catchment lithology and pedology. However, there is a strong relationship between inferred pH and the concentration of trace metals and carbonaceous particles found in lake sediments. A strong relationship also exists between the distribution of water subject to acidification and areas of acid deposition. These analyses indicate that acid deposition from fossil fuel burning has been the primary cause of acidification since 1800. Acid deposition can be dry, wet (rainfall) or occult (mists or fog). Concentra- tion of deposited suiphates and nitrates may occur due to evapotranspiration. Once in the soil, final water acidity is determined by total available buffering capacity.

A further accelerated pattern of acidification is evident in some lakes from 1930 to 1970, and particularly in Wales from around 1940. These observations have led to a land use hypothesis which suggests that upland acidification is associated with softwood afforestation and the cessation of liming. The impact of afforestation was noted for Welsh catchments, where low rates of acid deposition could not account for rates of acidification, despite the low available buffering capacities (Stoner 1985). Evidence of the impact of afforestation on acidification has come largely from paired catchment studies (such as Harriman and Morrison 1982; Stoner and Gee 1985; Reynolds et al. 1986) and from studies investigating water quality in rivers draining catchments with variable proportions of afforestation (Kay and Stoner 1988; Donald

204 Water quality issues in the British uplar~ds

1989). Long-term studies of acidification following afforestation, such as those at Llanbrynmair in Wales and Loch Dee in Scotland, do not show increases in acidity. However, evidence suggests that acidification will not take place until after canopy closure, some 15-20 years after planting (Warren 1989).

The most likely causal mechanism for acidification from afforestation is an increase in the efficiency of removal by needles of dry and occult pollutants (Reynolds et al. 1986; Gee and Stoner 1988). This would increase solute concentrations in throughfall and soil waters. Other possible mechanisms include concentration of solutes by increased evapotranspiration, removal of base cations by tree uptake, improved drainage preferentially leaching base salts, and oxidation of elemental sulphur and nitrogen to sulphate and nitrate (Stoner 1985).

Upland areas have traditionally been subject to lime application to improve grassland. The reduction in liming subsidy in 1959 and its subsequent removal in 1976 has led to a marked reduction in application rates. The impact of these reductions on acid-susceptible areas is not simple to assess, but acidification may be enhanced if the buffering capacity of surface waters falls below 2 mg 1-r (Ormerod and Edwards 198.5). Alternatively, enhanced acidification may follow the replacement of liming grants with improvement grants (Boon and Kay 1990) promoting the application of high-nitrate fertilizers which may increase acidification by the oxidation of ammonium ions (Brady 1974).

Future trends in acidification are dependent on changes in acid deposition and land use. Modelling allows the effect of changes in these factors on acidification to be predicted. Models exist which predict short-term changes in acidity through storm events, where the level of acidity is determined by the ability of the catchment to neutralize acid inputs during the hydrograph. These models include the empirically based CAPTAlN (Whitehead et a/. 1986), and the BIRKENES (Christopherson et a/. 1982, 1984), PULSE (Bergstrom etai. 1985) and ILWAS (Gherini etal. 1985) models. The latter three, which include varying complexities of hydrology and soil chemistry, are process based and are not therefore catchment specific. Models for the prediction of future long-term trends in acidity include RAINS (Kamari et a/. 1984), TRICKLE DOWN (Schnoor et al. 1984) and, perhaps the best known, MAGIC (Cosby et al. 1985). MAGIC has the most sophisticated hydrochemistry and has been applied to a wide range of catchments in the US, the UK and Scandinavia.

Acidification may be reduced by land management. Softwood plantations increase acidity locally by around 30 per cent and will add to acidification in the UK if a policy of upland planting continues. If plantations are felIed and not replaced, short-term increases in acidity due to nitrate output are replaced by long-term reductions in acidity. Acid inputs may be neutralized by liming of catchments and water bodies (see, for example, Underwood et a/. 1987). To combat acidification in this way requires repeated applications of lime, which can be costly and have adverse effects on the naturally acid upland ecosystem.

The only effective method of restoring waters to their pre-acidification status is to reduce acid deposition. Models predict that a reduction in acid deposition of about 90 per cent is necessary to achieve full restoration, while a reduction of 50 per cent of current levels is needed to prevent further deterioration. Such extensive emission reduction requires action across Europe and is not foreseeable at present.

Trace metal mobilization

Many of the adverse ecological effects of acidification detailed above are in fact a result of the mobilization of potentially toxic metals, especially aluminium (Al),

Gordon N. Mitchell 205

under acid conditions (Warren 1989). Aluminium levels > 0*2mg 1-t have been shown to be toxic to salmonids in Welsh catchments (Stoner et ai. 1984; Stoner and Gee 1985). Aluminium also has significance for public health as it has recently been linked with the occurrence of Alzheimer’s disease (Flaten 1986).

Aluminium forms complexes with inorganic species (such as fluoride, hydroxide and sulphate) and organic molecules such as humic substances (HS). In both soil and water, the complexation reactions are strongly pH dependent, with monomeric aluminium more labile under increasingly acidic conditions. The complexation by HS of inorganic aluminium is particularly important as it renders the aluminium non- toxic. A detailed model (CHAOS) predicting the interaction of HS with inorganic aluminium has been developed and successfully tested on waters from across Europe and North America (Tipping et a/. 1988a, 1988b). The model describes the competi- tive interactions of aluminium species, protons and Ca2+ with HS and so can be used to estimate how the chemistry of acid environments responds to changing conditions.

Historical records of aluminium in water are poor, but from constructed pH records and physiochemical models like CHAOS it can be inferred that acidification has resulted in an increase in water-borne aluminium. However, acid deposition is not entirely responsible for increased concentrations of aluminium in runoff. The proportion of catchment covered by softwood afforestation has been correlated with aluminium in runoff (Kay and Stoner 1988; Ormerod et al. 1989). Higher concentrations of aluminium, and also manganese, from afforested areas were first noted in central Scotland (Harriman and Morrison 1982) and are a consequence of catchment acidification.

The concentration of other potentially toxic metals, such as lead and cadmium, in water depends on their geochemical availability. Little data exist on the distribution of heavy metals in UK surface and groundwater, but evidence from Sweden suggests that acidification does result in an increase in heavy metals in runoff (Jacks et a/. 1984). Studies of Welsh waters indicate that heavy metal concentrations are dependent on discharge and pH (Neal et al. 1986).

The mobility of many metal species is increased under acidic conditions. Therefore the geographical distribution of unacceptably high concentrations of potentially toxic metals is closely tied to areas subject to acidification. Measures to control acidification, such as emission and land use controls, are thus likely to be effective in reducing levels of toxic metals. For example, using regression equations it is possible to suggest maximum percentage forest cover for catchments in different stream hardness categories so as to achieve set water quality objectives (Ormerod et al. 1989). Such a scheme is currently being evaluated by Welsh Water using a maximum acceptable aluminium concentration of 0.08 mg I-t. This value is accepted as the toxic threshold to salmonids, but the scheme could equally be operated on EC directive guide or maximum acceptable concentration (MAC) levels depending on objectives.

Discoloration

Water draining upland catchments with organic-rich soils can dissolve large quantities of organic matter (DOM), which can impart a colour to water similar to weak tea. While upland water has always been coloured in this way to some degree, recent evidence suggests that the intensity of the coIour is increasing (Edwards 1987). Particularly high colour levels were experienced in some areas following the droughts of 1976 and 1984.

Discoloured water presents a serious threat to the quality of potable water. Water exceeding the EC standard (l-5 aurn-’ at 400 nm) is readily observed by eye and


often gives rise to complaint. During 1986, 1.5 x IO6 households in the Yorkshire Water region received water in excess of the EC standard. The colour can be removed at treatment works but recent increases in colour have greatly increased operating costs. However, many treatment works receiving upland waters apply only filtration and chlorination and may not be capable of removing the high levels of colour now experienced. Therefore treatment works are being upgraded or replaced at high capital cost (Yorkshire Water alone, for example, is to spend &85 million by 1995). However, a more serious problem posed by discoloured water is the potential threat to public health. During chlorination naturally coloured waters can form trihalo- methanes (THMs), such as chloroform and other non-volatile halogenated com- pounds, which are mutagenic and are suspected carcinogens (Lancet 1981).

Discoloured water problems have been reported in the Yorkshire, Severn Trent, Northumbrian, North West, Welsh and Dumfries and Galloway water supply regions (Edwards et al. 1987). A seasonal pattern of discoloration exists, with peak colour flowis experienced during the autumn. Statistical modelling of water colour from the upper Nidd valley indicates that discoloration relates to large soil moisture deficits 3 and 14 months prior to the annual colour peak (Naden and McDonald 1989). This model could account for 68 * 8 per cent of the variation in colour using the Meteoro- logical Office Rainfall and Evaporation Calculation Scheme (MORECS) soil moisture deficit data. These observations suggest that highly coloured runoff is associated with water table lowering and elevated aerobic decomposition of catchment peats, followed by catchment flushing.

Consistent differences in colour have been found in waters draining apparently homogeneous, adjacent catchments (McDonald and Naden 1987). This distribution of coloured water depends on catchment characteristics (McDonald et al. 1990) and can be predicted for unmonitored catchments from information on catchment peat coverage, drainage intensity, slope and land use (Mitchell 1991).

Water is coloured by organic matter decomposition products, such as humic and fulvic acids. These are produced by the action of micro-organisms whose metabolism is greatly increased by aerobic conditions, optimum temperature and a high surface area of organic matter exposed. Aeration is increased by water table depression following natural drying, moorland drainage or burning and is greater for south facing slopes. Temperature increases occur during the warmer summer months and may be enhanced by surface albedo changes following burning or erosion. The density of void spaces in peat increases with depth but the average void size decreases. This results in a great increase in organic matter surface area available for microbiological decomposition when the water table is depressed. Mechanical changes to the peat surface, such as desiccation cracks, will also offer a larger surface area for decomposition (McDonald et a/. 1990; Mitchell 1991).

To release decomposition products into water, thereby colouring it, requires a change in the hydrological status of the catchment. Severely dried peats do not re-wet easily due to macropore collapse and the production of hydrophobic organics. This can result in a delay of several months between water table rise and peak colour output. However, laboratory studies indicate that once peat is saturated the rate of colour removal is directly proportional to the throughflow rate (McDonald et al. 1990). A further colour model ‘driven’ by meteorological data, is under development (McDonald et a!. 1990; Naden 1990). This synthesizes the process information and should therefore be applicable to a wider range of catchments.

Future trends in discoloration are uncertain. It has been suggested that the high variability in rainfall over the last 30 years, with dry summers and alternately very wet and very dry springs, has now decreased (Jones and Wigley 1988). If so, model

Gordon A’. ~i~ehe~i 207

predictions suggest a reduction in colour. However, following studies of the ‘green- house effect’, the currently favoured climate scenario is drier springs and summers and wetter autumns. Model predictions suggest that this scenario would produce a greater variability in colour, with particularly intense colour in autumn runoff waters (McDonald et ul. 1990).

Although future levels of colour are unknown, action can be taken now to reduce discoloration. Control measures include land and catchwater management. Land management controls are designed to reduce the colour of runoff by sensitive operation of drainage and burning practices, and by the promotion of anti-erosion practices. Two catchwater strategies have been developed. The first is designed to discard colourcd water from water gathering in a manner that reduces overali discoloration while minimizing loss of yield. The second also reduces overall colour while minimizing loss of yield by using an automatic gate to exploit the lag time between peak discharge and peak colour (McDonald el al. 1990). These strategies are currently under evaluation by Yorkshire Water.

Nutrient loss

One of the factors limiting agricultural productivity in the uplands is the availability of nutrients for plant growth. Due to the low intensity of agriculture, large quantities of nutrients, particularly nitrogen, have accumulated. However, the low soil temperature, acidity and anaerobic conditions confine plant growth to a short summer growing season when mineralization and plant nutrient uptake are maximized (Batey 1982). During the winter, NPK demand by plants is low and available nutrients can be lost to ground and surface water through leaching or sediment loss, particularly during periods of high discharge (Roberts 1985). Levels of nutrients in waters draining upland catchments have traditionally been low due to the tight nutrient budget. However, land use changes, specifically grassland improvement and afforestation, have raised concern over the level of nutrients now finding their way into water- courses.

There are several problems associated with nutrient enrichment of upland water. Gastric cancers and methaemoglobinaemia in infants can result from high levels of ingested nitrate (Shuval and Gruener 1972) and the World Health Organization (WHO) has therefore set a limit of Il.3 mg I- ’ N03-N for potable water. Nitrate removal from drinking water supplies requires additional treatment that is not generahy available at works treating upland waters. Enrichment can therefore lead to non-compliance with set standards. Nutrient enrichment of impounded water may result in eutrophication and the growth of algal blooms. This can reduce the amenity value of impoundments and greatly increase the cost of water treatment, particularly if microstraining is not sufficient to remove all the algae. Nitrates may also add to acidification by the oxidation of ammonium. Finally, elevated nutrient levels in upland waters reduce the efficiency with which lowland waters, containing higher levels of nutrients from agricultural runoff, are diluted.

EC subsidies to marginal upland farms encourage increased sheep stocking densities and pasture improvement. Improvement consists of drainage, liming and phosphate application in the first year, seeding and NPK application in the second year, followed by applications of N and P at 3-4-year intervals. Drainage increases N03-N in runoff through accelerated aerobic decomposition associated with water table lowering, while increased hydraulic conductivities result in rapid leaching of applied nutrients, particularly in the first year of improvement when the land lies fallow. There have been few studies of nutrient losses from upland catchments, but


exceedances of water quality directives related to pasture improvement have been noted for N03-N (Roberts et al. 1983) and for NH4-N (Kay and Stoner 1988).

Since 1924 conifer planting has averaged 18 133 ha yr I, with a total area of over 2.2 x lo6 ha (Nature Conservancy Council 1986). The majority of this afforestation has taken place in the uplands, particularly in Scotland. Elevated nutrient losses occur due to pre-afforestation drainage, phosphate applications, scavenging of atmos- pheric nitrogen by mature trees, and from increased soil and brushwood decomposition following clear felling (Roberts 1985). Levels of NO,-N in streamflow following felling were found to exceed continuously the 11.3 mg l- ’ WHO limit for a full two years where revegetation was not encouraged (Likens et al. 1970). In the Loch Ard forest, Scotland, losses of NPK following fertilization were 4 per cent, 15 per cent and 20 per cent respectively (Harriman 1978). Levels of N and K returned to normal after two to three years, but P remained high after three and a half years. These leaching rates were thought to be potentially beneficial for stream productivity but harmful in impoundments.

More than 260 upland reservoirs are thought to be at risk of suffering nutrient enrichment problems (Youngman and Lack 1981). It is not clear how many upland supplies are now subject to enrichment, but eutrophication of lochs and reservoirs in Scotland has been reported (Holden and Caines 1974; Cooke 1974; Greene and Taylor 1989).

The historical pattern of nutrient loss in upland catchments is poorly recorded. However, the pattern of nutrient loss is associated with land use changes and nutrient levels in excess of WHO and EC directives are increasingly evident. Nutrient losses to water courses can be ameliorated by sensitive catchment management. For example, NO,-N losses are reduced when narrow strips of unfelled trees are left along stream banks (O’Loughlin et al. 1980). Nutrient levels may also be reduced in streams by managing the watercourse to increase the natural uptake of nutrients by sediments (Wyer and Kay 1984). However, the long-term future trend in nutrient loss from upland catchments is likely to be dictated by the levels of subsidy and tax incentive leading to pasture improvement and afforestation.

Sedimentation

There are several problems associated with high concentrations of suspended sediment in upland waters. High sediment yields can destroy aquatic spawning grounds and food supplies (Cordone and Kelly 1961) and damage fish stocks (Stewart 1963; Graessner 1979). In addition, highly turbid waters are difficult to treat and can block water treatment filters, reducing their capacity to produce potable water. High sediment yields from upland catchments may also reduce reservoir capacities, particularly where drainage basins are large in relation to the reservoir.

Erosion rates from undisturbed moorland catchments are considered low when compared to lowland agricultural catchments and do not pose any serious threat to water quality (Burt and Oldham 1985). However, disturbance of upland catchments can result in high suspended sediment yields that do cause problems. Sediment yields increase durring thunderstorms and can be especially high after periods of drought (Burt et al. 1983). Moorland burns, especially when large and uncontrolled, can result in a reduction of infiltration capacity and an increase in surface runoff and overland flow and so promote erosion. The degree of erosion is partly dependent on the extent and severity of the burn (Kinako and Gimingham 1980).

The operations having the greatest impact on suspended sediment yields from upland catchments are pre-afforestation ditching and road construction. At the Cray


reservoir in the Brecon Beacons increased turbidity resulted in costs to the water authority of &364 000 over the 18 months following these operations in 1981 (Forestry Commission 1988). Drainage of a small upland catchment in the North West resulted in a 50-fold increase in sediment yield, which over the five-year study period was equivalent to a half-century of erosion at pre-drainage rates. Sediment yields were found to remain at about four times pre-drainage rates (Robinson and Blythe 1982). Long-term elevated sediment yields following drainage have also been observed in the southern Pennines (Burt et a/. 1983).

Bathymetric surveys of 41 reservoirs in the Pennines has demonstrated a high degree of spatial variability in the loss of capacity, ranging from 0 to 75 ~2 per cent per century, with an average loss of 12 per cent (Butcher et al. 1989, 1990). The highest capacity losses were found in small impoundments in large, peat-covered catchments and represent small absolute losses. The spatial distribution of capacity loss and erosion rate is attributed to the spatial distribution of areas prone to erosion, determined using topographical indices and satellite imagery. Highest area-specific yields were found for the more recently constructed reservoirs, suggesting that erosion is currently active in the southern Pennines. However, the main impact of sedimentation on water quality is clearly from episodic erosion rather than capacity loss (Burt and Oldham 1985).

The main problem with sedimentation is the acute pollution experienced during some discrete erosion events. These are particularly prevalent following moorland drainage, and especially pre-afforestation ditching. While the implementation of erosion control strategies (see Forestry Commission 1988) may be beneficial in reducing sedimentation, future trends in sedimentation are likely to be heavily dependent on the expansion of upland afforestation. No national strategic evaluation of sedimentation following increased afforestation has been attempted to date.

Conclusion

In this paper an attempt has been made to identify why particular upland water quality issues are perceived as problems, and to examine their extent, historical trends, causal mechanisms and possible control methods. However, it is recognized that these upland water quality issues cannot be considered in isolation from each other, and that water quality management is only one aspect of water resource management. Similarly, water resource management is only one of several potentially conflicting activities that are carried out in the British uplands.

Water quality research has been typified by site-specific, single-parameter analysis (Kay and McDonald 1982). While this approach has undoubtedly increased our understanding of individual processes operating in upland catchments, it has not allowed synergistic effects to be properly assessed. Therefore, it may be difficult to qualify, let alone quantify, the impact on one water quality parameter of changes in another. For example, strategies designed to reduce discoloration of water may also be effective in reducing the severity of acidic flushes and fish kills in rivers. However, colour reduction strategies reduce the availability of organic molecules in water that can bind aluminium, rendering it non-toxic. To resolve such problems requires an understanding of the synergistic effects operating in upland catchments upon which water resource policy can be used.

Modelling provides a way in which these problems may begin to be addressed. Process-based models exist, or are being developed, for each of the issues examined in this paper. Following sufficient testing of these models there is potential for integrating them and so identifying the simultaneous impact of environmental change


on several water quality parameters. This would assist managers in assessing the impact of environmental changes, such as acid deposition or climate change, on water quality. The coupling of such an integrated model package with digital terrain data would enable it to be tailored to specific catchments, allowing managers to assess the impact on water quality of, say, land use change, and thus assist in optimizing water quality. Such developments lie within the domain of the embryonic field of analysis- based geographical information systems.

Although models and integrated model packages may be extremely useful in upland water quality management it should be recognized that they can only assist in decision-making. Decisions should be made in the context of comprehensive strategic water planning. For example, upland afforestation has been heavily criticized for its deleterious effect on runoff quality, but it has also been noted that upland afforestation produces a much lower load of enteric bacteria than grassland, which can be considered an advantage (McDonald et al. 1984). However, current Forestry Com- mission investigations in the Welsh uplands suggest that up to 12 per cent of the UK’s sewage sludge could be disposed of as a fertilizer for coniferous saplings. If such a practice were to be adopted nationally, what would be the impact on the sanitary quality of water, the single most important criterion for potable water? In addition, what levels of heavy metals, such as zinc and cadmium found in sewage sludge, may be expected in runoff in the future, and could the basic upland treatment plants cope with them? Upland water quality can only be effectively managed when there are clear strategic aims, and where there is an assessment of the costs and benefits associated with alternative decisions. Traditional cost-benefit analysis is certainly flawed, but upland water quality cannot be managed without some economic assessment of alternative options.

In the rural context, conflicts of interest exist between economic and amenity value, between local and national need, and between individual and community (Davidson and Wibberley 1977). Water resource management is just one of several conflicting activities in the British uplands. The present approach to upland planning is one of discrete policy-making by major upland interests. What is needed for the long-term benefit of the uplands, and therefore water resource management, is an integrated approach where multiple-use solutions are generated within the framework of upland activities. Only then can objectives be set and met by land use activities that are currently in conflict. In the long term, solutions to upland water quality problems are not technical, but economic and political.

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Water quality issues in the British uplands