A REVIEW OF GUIDANCE AND CRITERIA FOR MANAGING RESERVOIRS AND ASSOCIATED RIVERINE ENVIRONMENTS TO BENEFIT FISH AND FISHERIES
by
Leandro E. Miranda
Mississippi Cooperative Fish and Wildlife Research Unit
Post Office Box 9691
Mississippi State, Mississippi, USA
Few river systems have escaped impoundment. There are about 60 000 reservoirs worldwide with a volume larger than 10x106m3, representing a total volume exceeding 6 500 km3 and a surface of about 400 000 km2. Over 2 800 of these reservoirs have a volume larger than 100x106m3 and account for about 95% of the 6 500 km3 combined reservoir volume. Construction of dams has been driven by economic needs, while ecological consequences have received less consideration. Construction of reservoirs is slowing down in industrialized countries, but elsewhere construction continues at a rapid pace.
Reservoirs provide significant contributions to global fisheries. The main challenges to maintaining and enhancing reservoir fisheries and associated social and economic benefits are fish habitat and environmental degradation, inadequate fish assemblages, inefficient harvesting systems, stakeholder conflicts, and insufficient institutional and political recognition. This review examines existing measures, guidelines, and criteria available to manage reservoirs and their releases for the benefit of fish and fisheries in the river basin. The review considers measures for managing the environment, fish, and associated fisheries within the reservoir, as well as upstream and downstream. Issues associated with estuaries are also considered, but briefly. Once technical criteria and guidance relevant to managing the environment, fish assemblages and fisheries are addressed, the focus turns to improving management. Attention is given to administrative and procedural matters germane to the management process.
An important barrier to reservoir fisheries development and management is that fishery administrators find it difficult to defend the interests of their sector. Decisions over developments affecting fisheries and aquatic environments are often made with minimum or no consideration of these sectors, mainly for lack of reliable economic valuation and lack of political clout by the users. Given this lack of political power, the interests and needs of fishers and fisheries managers are often not properly represented within existing political frameworks, and thus neglected or ignored. Fishery administrators and stakeholders should seek every opportunity to communicate their needs, demonstrate the value of fisheries and the aquatic natural resource integrated by fish, and participate in the political process.
Management of impounded river basins in many developing countries follows models developed in North America or Europe. Strategies are often imposed by foreign experts or copied, considering neither climatic, faunal, socio-economic conditions, nor political realities. Despite the apparent commonality in environmental issues, management policy must be country-specific, and take local conditions into account; blind application of imported principles leads to policy failures. Although experience in other regions should not be ignored and should serve as the base for management plans, evaluation of strategies and revision to fit local realities are critical to successful adaptation and implementation.
A holistic approach to the management of fresh water as a finite and vulnerable resource must be taken, one that integrates economic, social, and environmental needs. Water allocation plans for river basins are essential to ensure that available water is adequately apportioned to meet this goal. The multi-sectoral nature of water resources development in the context of socio-economic development must be recognized, as well as the multi-interest use of water resources for water supply and sanitation, agriculture, industry, urban development, hydropower generation, inland fisheries, transportation, recreation, floodplain management and other activities. To this end, effective water management plans, coordination, and implementation mechanisms must be in place in all river basins.
Resource agencies can increase their effectiveness in developing reservoir mitigation by fully participating in the development process. However, to effectively recommend mitigation procedures, agencies need to incorporate technical expertise in fields other than traditional fish ecology and management, coordinate among other agencies, be willing to make recommendations based on incomplete information, have procedural expertise, and develop effective policies. Clear and effective policies for reservoir development can enhance an agency's influence in developing mitigation. The more thoroughly an agency can back-up mitigation recommendations with established regulations, policies, and specific scientific objectives, the more influence the recommendation will have. Broad policy and goals must be transformed into clearly defined targets and objectives.
Licensing and relicensing of dams may be used to ensure that reservoir construction and operation weights environmental concerns. Then, when deciding whether to issue or reissue a license, conservation, protection of fish and wildlife, fisheries, recreational opportunities, and preservation of general environmental quality benefits can receive equal consideration to energy or other economic benefits provided by impoundment. A relicensing process would provide an opportunity to modify old dams, and address problems created decades before environmental issues moved to the forefront. Relicensing would also provide an important medium by which the evolving public interest about river conservation can be addressed.
Sound management of impounded rivers depends on an ability to understand the effects of natural and human-induced change, which make management of impounded river basins extremely complex. Properly designed monitoring programmes that include repeated observations over time can separate natural effects from human ones, and distinguish effective management practices from less effective or harmful ones. Monitoring programmes are needed to support a comprehensive, scientifically-based evaluation of the present and future condition of the environment and its ability to sustain present and future populations.
Our understanding of impounded rivers and ability to predict how they will respond to management actions is limited. Together with changing social and economic values, these knowledge gaps lead to uncertainty over how best to manage impounded rivers. Despite these uncertainties, managers must make decisions and implement plans. Adaptive management is a way for managers to proceed responsibly in the face of such uncertainty. It provides a sound alternative to either charging ahead blindly or being paralysed by indecision, both of which can foreclose management options, and have social, economic and ecological impacts.
Impounded rivers are relatively new aquatic ecosystems in the global landscape. They represent an important economic and environmental resource that provide benefits such as flood-control, hydropower generation, navigation, water supply, commercial and recreational fishing, and various other recreational values, particularly in developed countries. The ecological impacts of impounding a river have been dramatic and extensive. Construction of dams has been driven by economic needs, while ecological consequences have received less consideration. Construction of reservoirs in industrialized countries has slowed down after a peak near the middle of last century, because nearly all suitable sites have been impounded and ecological concerns have become more prevalent. In other parts of the globe, construction continues at a rapid pace (Avakyan and Iakovleva, 1998). Consequently, criteria to develop new reservoirs and understand their impact remain a need in some parts of the world, whereas demand for criteria to manage existing reservoirs to maximize their benefits and minimize or mitigate their ecological impacts continue to increase.
The geographic distribution of reservoirs reflects a complex interaction between topography, climate, and the economic need to control water movement through river basins. Water management goals tempered by climate have guided reservoir design, whereas economics tempered by topography have dictated their construction and operation. Some large river systems have been transformed into cascades of reservoirs by stacking them in chains. Examples include the Paraná in South America; Arkansas, Columbia, Missouri, and Tennessee in North America; Volga and Dnieper in Europe; Angara in Asia; Zambezi in Africa; and a many other large and small rivers. There are about 60 000 reservoirs with a volume larger than 10x106m3, representing a total volume exceeding 6 500 km3 and a surface of about 400 000 km2 (Avakyan and Iakovleva, 1998). Over 2 800 of these reservoirs have a volume larger than 100x106m3 and account for about 95% of the 6 500 km3 combined reservoir volume. Of these larger reservoirs, for which records are more reliable, 915 (1 690 km3) occur in North America, 265 (971 km3) in Central and South America, 576 (645 km3) in Europe, 815 (1 980 km3) in Asia, 176 (1 000 km3) in Africa, and 89 (95 km3) in Australia and New Zealand (Avakyan and Iakovleva, 1998). Few river systems have escaped impoundment.
Reservoirs provide significant contributions to global fisheries (Moreau and DeSilva, 1991; Petrere, 1996; Fernando et al., 1998; Miranda, 1999). In North America and Europe, and more recently Australia, recreational fisheries are economically important, and the trend is for increased importance elsewhere. For instance in the USA, reservoirs attracted about 21 million recreational anglers in 1990, and supported over half of all freshwater fishing (USFWS, 1993). Commercial fisheries are most important in Asia, but are also important in Africa and South America. In some parts of the world, reservoir commercial fisheries are essential for subsistence and often represent an irreplaceable source of high-quality and low-cost animal protein crucial to the balance of human diets. Fish harvested from reservoirs are generally marketed regionally, and contribute to the livelihood of impoverished people and local economies. The main challenges to maintaining and enhancing reservoir fisheries and associated social and economic benefits are fish habitat and environmental degradation, inadequate fish assemblages, inefficient harvesting systems, stakeholder conflicts, and insufficient institutional and political recognition.
The purpose of this review is to examine existing measures, guidelines, and criteria available to manage reservoirs and their releases for the benefit of fish and fisheries in the river basin. Section 2 examines measures for managing the reservoir environment, as well as those upstream and downstream that influence the reservoir, or are influenced by the reservoir. Section 3 considers measures for managing fish stocks and fisheries in the reservoir and associated riverine environments. Section 4 identifies selected avenues for improving guidance, criteria, and the processes of managing impounded rivers.
Point and non-point source pollutants imported from the watershed affect environmental quality in reservoirs. Point source pollutants consist mainly of contaminants and particulate and dissolved organic matter that can be traced to a localized source (e.g. a sewage outlet, industrial discharge). Non-point pollutants originate from diffuse sites (e.g. farm lands) throughout the watershed and include silt and clay, inorganic nutrients, particulate and dissolved organic matter, and contaminants (e.g. pesticides). The reservoir sediments represent a composite of materials originating from point and non-point sources. Although point sources can often be regulated, non-point sources in large watersheds are more difficult to control. Nevertheless, interagency efforts that result in watershed and reservoir improvements have been documented (Born et al., 1973). Sediments, excessive nutrients, organic matter, and contaminants may be kept out of reservoirs through proper agricultural and waste management practices.
Reservoirs effectively trap suspended solids, sedimentation increases turbidity to limit primary production, and decreases depth and thereby storage capacity, all of which affect various physical and chemical processes that eventually influence the biotic community. Sediments originate from erosion processes within the drainage area, the river channel and the reservoir shore, and are normally the major non-point source of pollution. Sedimentation is aggravated in basins characterized by prolonged droughts followed by seasons with heavy rains such as in Brazil (Gomes and Miranda, in press, a), or by monsoon seasons such as in India (Sugunan, 1995). Practices to reduce loads of sediments and nutrients entering rivers and reservoirs include conservation tillage, terracing, crop rotation, vegetative cover, crop residue, nutrient management, streamside management zones and use of structural devices such as retention basins, sediment dikes, and erosion control weirs (Johengen et al., 1989).
Phosphorus inputs and regeneration from sediments are the major factors driving autotrophic production. Forsberg & Ryding (1980) considered lakes with total phosphorus concentrations <15µg L-1 as oligotrophic, 15-25 as mesotrophic, 26-100 as eutrophic, and >100µg L-1 as hypereutrophic. Phosphorus inputs often come from agricultural sources and municipal effluents. Excessive inputs result in rapid eutrophication. The consequences of eutrophication are algal blooms, which cause decreased water clarity, wide dissolved oxygen fluctuations, and dense littoral beds of aquatic vegetation. Through effects on water physical and chemical characteristics, dense algae blooms make reservoir environments unsuitable for many fish, reducing diversity of fish assemblages and fisheries. Management strategies to curtail eutrophication include dilution and flushing, aeration of hypolimnion and circulation to encourage utilisation of nutrients, precipitation and inactivation, sediment removal, water level drawdown, hypolimnetic discharges, and harvesting of biota (e.g. macrophytes and fish) (Cooke et al., 1993).
Inputs of organic matter can be beneficial or harmful, depending on the natural fertility of the basin and the nature of the system. Particulate organic matter is the basis of food chains in large, turbid reservoirs where autotrophic production is limited by turbidity and low retention time. However, smaller reservoirs that tend to be higher in watersheds are often less turbid, have higher retention, and are more autotrophic, and can thus be overwhelmed by the oxygen removal due to decomposition of large organic inputs.
Contaminants may include metals, pesticides, oils, and other pollutants in industrial, agricultural, and urban wastes. Elevated levels of these substances can reduce fish survival, reproduction, and growth and may bioaccumulate in fish tissue often rendering them unsuitable for human consumption (Cairns et al., 1984). Possible corrective actions include eliminating the point sources and managing the watershed (Baker et al., 1993). Aeration of the hypolimnion can oxidize some contaminants and encourage precipitation. Changes in water level may need to be adjusted if the exposure and suspension of contaminated sediments tend to increase the solubility and mobilisation of contaminants. Removal of contaminants is sometimes not possible and no action may often be the only approach, allowing natural processes such as sedimentation to gradually remove the substances. Such an approach requires considerable monitoring and possible closure of some fisheries.
Sedimentation problems on regulated rivers are considerably different from natural rivers. Problems include sedimentation in reservoirs, channel effects downstream of dams, and in-stream flow requirements for maintenance of channels and fisheries. Reservoirs can profoundly affect the geomorphology of streams that have a large natural sediment load, as the reservoir traps sediments and releases clear water. The resulting downstream geomorphic effects of clear water releases from dams typically include channel instability (as the channel and banks are eroded to satisfy the sediment carrying capacity of the waters), channel bed armoring, and alteration of habitat (Collier et al., 1996).
Instability of tributary streams can produce negative effects on habitat for riverine fish and those that occupy the reservoir but need riverine habitat to complete their life cycle. The extent and magnitude of stream bank erosion is greatly increased by activities that remove riparian vegetation, increase water flows, increase sediment movement, or otherwise cause channel down-cutting (Meehan, 1991). Large vessel traffic and the resulting wakes also can create bank scour. Moreover, in-stream mining can cause multiple effects on fish habitat.
Attempts to deal with bank erosion often involve the use of hard materials (Orth & White, 1993). In smaller streams, particularly those that seasonally become dry or nearly dry, bulldozing of streambed gravel against the banks has been a common practice to retard erosion. In rivers, the placement of rip-rap rock, broken concrete, and mixtures of materials (ie rocks, soil, branches) along the banks has been a common practice. More recently, soil and vegetation bioengineering such as grading or terracing a problem stream bank or eroding area, and using interwoven native vegetation mats installed alone or in combination with structural measures, are being used to stabilize stream banks. Once these features are protected, they can also serve as a filter for surface water runoff from upland areas, and as a sink for nutrients, contaminants, or sediments.
However, erosion of banks is a natural process of rivers. River channels move freely from side-to-side across floodplains by eroding the banks on one side while depositing sediment on the other. This dynamic characteristic results in constant creation of new fish habitat and many species depend upon a constantly changing river. Impoundments have changed how the river erodes its banks, and in many cases the erosion is occurring through down-cutting, but new lateral habitat is not being created.
In reservoirs with long stretches of riverine environment upstream, large lateral tributaries, or both, native potamodromous fish comprise a large proportion of the fisheries (Agostinho et al., 1995). These reservoirs often have low retention times and are generally not conducive to development of lacustrine faunas (Gomes & Miranda, in press, a). In such reservoirs, upstream riverine habitats should be a conservation priority. Examples include Itaipu Reservoir, where a 230 km stretch of the free-flowing Paraná River connected to an extensive floodplain provides important spawning and rearing habitats for many species in the fishery. In 1998, at the urging of conservationists and fishers, this section of the river was set aside as a national park by the Brazilian government, to prevent further impoundment, degradation, and enhance fisheries in Itaipu Reservoir (Agostinho et al., in press).
Characteristics of a river along a continuum from its upper to lower reaches often change continuously and dramatically. Characteristics that may be greatly modified by a dam such as temperature, nutrient levels, plankton production, and biodiversity may be affected differently by damming the upper, middle, or lower reaches of a river system (Ward & Stanford, 1983). A headwater dam can greatly depress the ratio of coarse particulate to fine particulate organic matter below the impoundment because in-stream transport of detritus is blocked, whereas impounding the lower reaches of the river may have little effect on the composition of detritus. Correspondingly, functional feeding groups of invertebrates and fish can reflect these changes in detritus, whereas a dam on the lower reaches may not greatly alter the trophic relations of the receiving stream.
Stream characteristics are sometimes most diverse in the middle reaches. For example, flow discharge patterns in the upper and lower reaches of rivers tend to be less variable. In the upper reaches, constancy of discharges may be influenced by feeding springs and by the moderation of precipitation by terrestrial watershed processes; in lower reaches, excessive variation may be prevented by the collection of waters from many tributaries (Hynes, 1970). The middle reaches are generally most influenced by local meteorological events that may vary widely among geographical regions within the river basin, and hence these reaches tend to exhibit the most variable and unpredictable flows. Reservoirs moderate flow fluctuations in the middle reaches by storing water during periods of major runoff and releasing water during period of low flow (Ward & Stanford, 1995). Thus, reservoirs can have more impact on flow regimes of middle reaches as compared to upper and lower reaches.
These patterns of environmental heterogeneity along a stream continuum influence biodiversity. The high biodiversity in the middle reaches of a river system may result from spatial and temporal heterogeneity (ie unpredictability of flow, temperature, and other characteristics); increased predictability due to impoundment may result in reduced biodiversity. Impoundments in the upper reaches, however, are likely to influence biodiversity less by altering environmental heterogeneity and more by disrupting transport and processing of allochtonous materials. Impoundments in the lower reaches can limit migratory fish. In general, effects of impoundments along a river system continuum are likely to be less severe when dams are placed in the upper reaches than when placed in middle or lower reaches; nevertheless, severity of effects is related to a multiplicity of other factors that may easily override the effect of location in the river continuum.
Moreover, reservoirs positioned in the upper reaches have higher water retention times, and thereby develop fish assemblages more characteristic of lake environments. In the Paraná basin, large drainage areas and low water retention times of reservoirs in the lower reaches, exacerbated by high rainfall during spawning and rearing periods, limited the emergence of fit reservoir species from within the existing pool of riverine species (Gomes and Miranda, in press, b). The resulting assemblages have characteristics that are neither riverine nor lacustrine, and are maladapted to support fisheries. Introduction of lacustrine species (discussed below) are destined to failure because environmental characteristics are not lacustrine, except in reservoirs positioned high in the watershed where increased retention time allows development of lacustrine, autotrophic conditions. For example, Billings Reservoir, high in the Upper Paraná basin, exhibited greater hydraulic retention time and more autotrophic production, characteristics of lacustrine environments where assemblages rely more on grazing food chains. Coincidentally, this reservoir supported the greatest fisheries yield in the basin.
Various considerations are juggled in siting a reservoir. Engineering economics dictate that reservoirs be constructed with a high ratio of reservoir volume to dam volume, and thus reservoirs are preferably built in narrow, steeply sloped reaches that have broad valleys branching upstream. Other site requirements include topographic relief that provides a favorable reservoir surface area to volume ratio (ie sufficient storage volume without excessive shallow areas); and an alignment that does not allow a strong influence of prevailing winds (to avoid erosion and excessive turbidity). From an environmental perspective, biodiversity of the riverine fauna is a major philosophical issue. Impoundment of a river system will generally negatively impact biodiversity; thus, rivers with unique faunas may need to be spared. Effect of human populations is another major concern. Siting a reservoir near a populated area may enhance quality of life through economic and technological advances, while providing various recreational opportunities. Conversely, such siting may lead to relocations or severe cultural impacts. Prognosis for eutrophication should be considered; rapid eutrophication leads to decreased return on investments and short-lived reservoirs.
Some physical features of the reservoir basin can be avoided or modified during construction to substantially improve fish habitat and fisheries. There are various specific factors relevant to fisheries management that should be considered during site selection and reservoir design. The watershed should be well vegetated to prevent excessive erosion that leads to sedimentation, and free of pesticides and other pollutants. Runoff from agriculture, livestock operations, and industrial sites should be treated or diverted. Limited shore development and no septic discharges should be allowed. Watershed size influences water level fluctuation and retention time. The optimal watershed size depends on lake volume, rainfall, topography and land uses. Retention times less than about four weeks limit autotrophic production and retention times larger than about a year are associated with strong stratification (Søballe and Kimmel, 1987; Søballe et al., 1992; Stra_skraba, 1999). The level of vegetation removal from the reservoir's basin during construction varies depending of the main purpose of the reservoir. It is often desirable to leave blocks of timber and brush in embayments and shallow arms to enhance fish habitat. Although most of the evidence comes from the west, DeSilva (1988) corroborated such practice in Asia but called for more research. Often fish cover and spawning habitat can be improved during reservoir construction. Extremely deep reservoirs are generally unproductive, whereas extremely shallow reservoirs may remain turbid or support excessive growth of aquatic macrophytes. Additionally, the minimum flow requirements to support aquatic organisms in downstream areas should be addressed. Determination of minimum flows is often problematic. Minimum flow requirements are typically determined to protect or enhance one or a few harvestable species of fish; other fish, aquatic organisms, and riparian wildlife are assumed to be protected by these flows as well.
Wetting through water level fluctuations saturate previously unsaturated material, potentially resulting in massive slides when the water level is drawn down (USACOE, 1987). This material accumulates at the base of the slope, and often forms an underwater bench, leaving steep unstable slopes above the water line. Reservoir banks are also subjected to attack by both wind itself and waves generated by wind or power boats, which tend to remove this material and undercut the banks. Sediment deposits originating from erosion within or outside the reservoir are of concern, as they impact not only the reservoir storage capacity but also many biological processes. Sediments deposit not only in the lower reservoir zone frequently reserved as a sediment pool, but often throughout the littoral causing shallowing and loss of productive substrates. The suspended solids and sediments derived from erosion and deposition damage fish habitat and produce adverse impacts like those described for non-point sources of sediments. Barren shores and mudflats are poor food producers, unsuitable habitat for nest builders, and poor nursery habitat for fish (Meals and Miranda, 1991).
Some water level management practices can minimize erosion and sediment deposition. Minimisation of the rate with which water level is drawn down helps reduce slides. Keeping reservoir level as low as possible during known periods of high sediment inflow encourages sediments to deposit in the lower zones of the pool. Periodically raising water levels high enough to inundate existing sediment deposits, reduces the establishment of permanent vegetation and subsequent increased sediment trapping in backwater reaches (USACOE, 1987).
Landscaping and waterscaping options to reduce shore erosion and sediment deposition include steep sloping shores with terracing above the water line; construction of retaining walls, rip-rapping with rocks, and installation of gabions; construction of breakwater structures to reduce the energy of waves; and encouraging growth of aquatic or terrestrial vegetation nearshore in reservoirs not subject to large water level changes. The most sensitive sites are often those exposed to long fetch. Buffer strips of natural vegetation along the shore also help to stabilize shorelines, reduce sediment inputs, and provide shading that produces cooler water temperatures. Protection against shore erosion can be very costly, and often must be limited to specific areas of concern (Summerfelt, 1993).
The effects of shore erosion on the aesthetics of the reservoir are closely related to its use for recreational purposes, and should be considered in the management of water level fluctuations. Often, aesthetics may be improved simply by minimising the duration of exposure of unsightly shorelines resulting from erosion. However, in some reservoirs it may be virtually impossible to compensate fully for such effects and still maintain the integrity of the functions for which the reservoir was constructed.
Fish and fisheries generally benefit from availability of moderate levels of plant cover; thus, absence of aquatic vegetation can be as harmful as excess (Dibble et al., 1996). Aquatic macrophytes also help stabilize sediments and shoreline, reducing problems associated with erosion and turbidity. The goal of aquatic plant management is to identify and provide an appropriate level of aquatic vegetation taking into account the effect of plants on the existing fish assemblage, reservoir physical characteristics, and uses of the reservoir.
Several management measures are available for reducing excess aquatic plants (Baker et al., 1993; Summerfelt, 1993). The control technique to be used is dependent upon the species of aquatic plant causing the problem, magnitude of infestation, distribution within the reservoir, and characteristics of the reservoir, and full consideration should be given to the impact of the control measures on fish and fish food organisms associated with aquatic plants. The degree of control required to bring the problem to an acceptable level must also be a consideration. Aquatic plants that cause problems in reservoirs are generally of the floating and submersed types. Biological, chemical, and mechanical methods, individually or in combination, may be used to control aquatic plants. Biological control employs organisms that feed on the target organism or affect it in some other way to reduce its density or growth. Biological control agents potentially available for use in aquatic plant control are insects, plant pathogens, and herbivorous fish. The application of safe and effective chemical agents is a proven method for aquatic plant control. Approved chemical agents for aquatic use may be liquids that can be sprayed onto floating plants or inserted under the water for controlling submersed plants, or they may be solids that can be applied by spreaders. Chemical methods are generally readily available and are relatively inexpensive when compared to other methods, unless used in large areas. Mechanical devices for controlling aquatic plants vary from deflecting booms and screens or clipping bars mounted on boats, to more sophisticated systems whereby the plants are cut and removed from the water to disposal areas. Although most mechanical methods are generally rather costly, they are sometimes desired over other methods because no organisms or chemicals are added to the environment. In reservoirs where water level can be lowered, reductions of some submersed species may be achieved through exposure to drying or freezing; however, some species are not affected or even encouraged by such drawdowns. In general, biomanipulation through introduction of herbivorous fish is the less costly control method where infestations are large. Nevertheless, introductions may be environmentally costly if fish leave the reservoir and affect desired aquatic vegetation in streams, associated wetlands, or even estuaries (Summerfelt, 1993).
Establishment of aquatic plants has proven difficult in those reservoirs that do not already have them, and some reservoirs may never develop an aquatic plant community. The lack of aquatic vegetation in many reservoirs may be attributed to the absence of propagules in the impounded basin, but may also be due to inhospitable environments for establishing seedlings (e.g. turbidity, unnaturally fluctuating water levels, inadequate substrates, excessive density of herbivores). Through time, seed banks may develop and reservoirs develop aquatic flora. Through planting, management of water levels and turbidity, and exclusion of predators, aquatic plants can be established and the system driven towards a vegetated, clear state (Smart et al., 1996).
Low temperature and concentrations of oxygen tend to occur in the hypolimnion when reservoirs stratify. Vertical stratification occurs mainly through the interaction of wind and temperature and creates density gradients that affect water quality. Intensity of the gradient varies latitudinally, and is affected by physical characteristics such as reservoir depth, water retention time, and water level fluctuation. Oxygen stratification is undesirable because anoxic conditions in the hypolimnion limit habitat availability and can impact water quality throughout the reservoir and downstream.
Methods to destratify or prevent stratification include hypolimnetic discharges, air bubbling/injection to generate water movement, and mechanical pumping between the hypilimnion and epilimnion to either generate water movement, or to aerate hypolimnetic water by passing through baffle systems (Ruane et al., 1986). A bubble column produced with compressed air will create upwelling in a reservoir that, in combination with wind energy, can be used to prevent stratification or to destratify (Pastorok et al., 1982). Mechanical pumping can also be used to avoid oxygen stratification without disrupting temperature stratification, by lifting hypolimnetic water to the surface where gases such as methane, hydrogen sulfide, and carbon dioxide are dispersed, and then water is returned to the hypolimnion without substantial increases in temperature (Wirth, 1981), which may allow maintaining both warmwater and coldwater fish communities and fisheries. Aeration of the hypolimnion through injection of oxygen has been reported to be more cost effective than through lift systems (Mauldin et al., 1988). Potential benefits of artificial destratification include expanded habitat for invertebrates and fish, improved quality of water stored and released, retarded eutrophication, and avoidance of catastrophic turnovers.
The littorals of reservoirs are usually highly unstable when water levels fluctuate. Fluctuations may be daily, seasonal, or annual depending greatly on the reservoir purpose, and present many problems to fish habitat management. However, operation curves often have enough flexibility to allow development of management opportunities.
Water levels have a direct effect on benthos, periphyton, and aquatic macrophytes abundance, but only indirect effects on phytoplankton and zooplankton (Ploskey, 1986). Increases in water level that inundate lush vegetation temporarily increase supply of food and cover, whereas extensive drawdowns concentrate fish and can increase foraging efficiency of predators. Reproduction of littoral spawners may be encouraged by flooding of shores (Miranda et al., 1984), but can be adversely affected if levels increase or decrease rapidly during the reproductive period, particularly nest builders or substrate spawners. Because year-class strength of fish varies annually depending on environmental conditions, optimisation of water levels for fish spawning may not be essential or productive every year.
Because substrates exposed by drawdown are subject to erosion, establishment of herbaceous vegetation after drawdown is important for erosion control and aesthetics. To be effective, drawdowns must occur during the growing season for successive seeding of herbaceous terrestrial vegetation and should allow for substantial grow before winter. Such drawdown also provides the opportunity to improve submerged structures that enhance habitat diversity for littoral species. Managers in temperate North America generally drawdown reservoirs in late summer or fall to allow establishment of terrestrial vegetation naturally or by seeding, flood terrestrial vegetation in spring when most species spawn, maintain constant water level during spawning season, and maintain water level as high as possible until the following drawdown (Willis, 1986). In general, water level manipulations are most effective when they are extensive, last several months, occur during the growing season, and flood or drain productive areas. On very large reservoirs where drawdown even of a small magnitude would result in huge water losses, this method of optimising littoral conditions for fish may not be practical.
In some African reservoirs water level must be managed to maintain intermediate levels of submersed aquatic vegetation. The aquatic vegetation provides food and shelter for key species in the fisheries. Bernacsek (1984) proposed to fluctuate water level 2.5 to 4 m annually to disallow excessive proliferation without eliminating the vegetation, which is possible with greater fluctuations. Drawdown rates not to exceed 0.6 m per month were needed to allow adjustment of littoral communities to the fluctuating water levels. In reservoirs of the Tennessee River, USA, the Tennessee Valley Authority stabilises water level within + 0.6 m for at least two weeks once water temperature reaches levels at which most littoral fish species begin spawning.
Where a system of reservoirs exists, operational flexibility to provide appropriate seasonal water level increases further down the chain. With coordinated scheduling, flooding and drawdowns may be provided to a reservoir at least every 2-4 years. Power and water demands may be met by releases from other reservoirs in the chain. Fishery managers should be familiar with rule curves (man-made hydrographs), operational needs, and fishery needs to be in a position to suggest modifications to water managers.
Fish production in some reservoirs may be limited by the availability of suitable spawning sites or poor quality of available sites (Summerfelt, 1993). Siltation is a major cause of degraded spawning habitat, but silt removal is impractical in most situations. Instead, raising water level during the spawning season, annually or otherwise, often creates suitable spawning substrate for nest-building species. In some instances it may be possible to construct sites where important species may spawn (Karpova et al., 1996). Gravel beds and reefs constructed of large rocks often attract reservoir species. Some species spawn in coves, lagoons, flooded wetlands, or other heavily vegetated flooded areas in river floodplains. These areas may also serve as nursery areas. These areas are often present in the headwaters of the reservoir and should be protected and managed; alternatively, raising water level during spawning and rearing period may artificially produce similar environments.
Some reservoirs may lack sufficient structural features to provide shelter. Without adequate shelter, survival of young fish of many species is often low, preventing adequate fishery yields. Structural features provide safety from predators, substrate for food organisms, and even spawning habitat. Also, some predators tend to concentrate around structural features searching for prey, and such concentration may enhance fishing success. Common types of physical structures include drowned or fallen trees, brush piles, rock reefs, and sharp changes in the bottom topography. Reviews on addition of physical structures to provide cover and spawning areas are provided by Ploskey (1985) and Brown (1986).
Aging of reservoirs is intricately linked to inputs from the watershed, and siltation is perhaps the most dominant aging process. Siltation reduces depth, affecting storage capacity of the reservoir and most importantly the characteristics of littoral habitats, particularly in embayments. These long-term depth reductions lead to slow changes in bottom firmness, as well as average, minima, and maxima values of temperature, oxygen, and other vital water quality conditions. Silt is likely to be rich in nutrients and organic matter also imported from the drainage area, which become available for primary production. Ultimately, this production promotes further release of organics and nutrients as they decay, and further eutrophication of the sediments. Reductions in oxygen prompted by decreased depth are exacerbated by increased sediment oxygen demands. In arid, agricultural regions of the dry zone of Asia, aging is also associated with salinisation prompted by excessive evaporation and leaching of minerals in soils (Petr and Mitrofanov, 1998).
As the reservoir ages and siltation progresses, nutrient dynamics of the system begin to change. With increased nourishment, phytoplankton communities shift from a domination by green algae to blue-green algae. Although dominance may also shift seasonally, in highly eutrophic reservoirs blue-green algae tend to dominate for an increasingly larger portion of the year (Wetzel, 1983). Zooplankton composition is affected by phytoplankton availability. Macrofiltrators (usually large-bodied zooplankton) are more abundant in younger, oligotrophic reservoirs, sometimes giving way to low-efficiency, small-bodied, algal and bacterial feeders as reservoir age and nutrients increase (Taylor and Carter, 1998). Additionally, in highly eutrophic aging reservoirs the food supply of zooplankton may actually decrease because of the dominance by blue-green algae, which are mostly inedible due to their large size (Porter, 1977). These interrelations between phytoplankton and zooplankton can have repercussions higher through the food web. As a result the fish community shifts to a less desirable assemblage of increased benthophagous and phytophagous species and reduced predator densities. Nevertheless, these changes are likely to occur over long terms.
Increased nutrient loads encourage growth of free-floating macrophytes, such as Eichhornia spp., whereas depth losses foster growth of rooted aquatic macrophytes and expand their distribution from shore (Cooke et al., 1993), if water level fluctuations are not large. Extensive macrophyte development can exert enormous control over the aquatic ecosystem, beginning with the physical and chemical characteristics of water (e.g. temperature, light, oxygen). Once extensive covers of submersed aquatic macrophytes (e.g. Hydrilla, Myriophyllum) become established, they can aggravate eutrophic conditions of an aging reservoir through growth-death-decay cycles that allow release of nutrients trapped in sediments, under anaerobic or aerobic conditions.
Other major change induced by aging is the deterioration of various habitats, particularly in the littoral zone (Benson, 1982). Standing timber often left in the basin decompose and fall. Long-term bank erosion induced by wind damage of exposed shorelines turns diverse shoreline habitats into uniform, barren mudflats. In reservoirs with substantial water level fluctuations, this effect is not limited to the normal pool shore, but extends into areas above and below normal pool elevation, and can expand a substantial percentage of the littoral area depending on the slope of the reservoir basin. The original productivity of this ecotone is thereby lost with age, and its instability precludes colonisation by terrestrial or aquatic flora.
Elimination or reduction of spawning grounds, or delayed access to the spawning areas have been the most significant effects of physical barriers. Dams can affect fish by blocking upstream and downstream passage. These movements are most important to anadromous and catadromous fish, which spend part of their life cycles in rivers and part in oceans or other large waterbodies, but also to potamodromous species which, during a certain phase of their life cycle, depend on longitudinal movements within the river system (FAO, 1998). For fish trying to move upstream, a dam can pose an impassable barrier unless passage is provided, and fish moving downstream are at high risk of being entrained in the turbine intake and injured or killed during downstream passage.
The blockage of upstream fish movements by dams may have serious impacts on species whose life history includes migrations for various purposes, e.g. spawning, feeding. Anadromous, catadromous, potamodromous, and some resident fish could all have spawning migrations constrained by such barriers. Maintenance or enhancement of such species may require the construction of facilities to allow for upstream fish passage. Descriptions of the design and functioning of various types of upstream passage facilities have been provided by many authors (e.g. Orsborn, 1987; Larinier et al., 1994; Clay, 1995; DVWK, 1996; Jungwirth et al., 1998). Upstream passage facilities can broadly be divided into two general categories (FAO, 1998): largely natural structures (e.g. fish slopes or rock ramps and by-pass channels) and more technical ones (e.g. pool-type fish passes, Denil fish passes, vertical slot passes, fish lifts and fish locks). Trapping and hauling is yet another option for bringing fish upstream around an obstacle. Fish passes are widely used to allow fish to get over single obstacles such as dams and may also be used to collect fish for hauling to upstream releasing locations.
According to Orsborn (1987), the term "fishway" describes any flow passage that fish negotiate by swimming or leaping, e.g. an artificial structure such as a culvert, a series of low walls across a channel (weir-and-pool fishway), or merely a chute up which the fish swim. Additional definitions are given by Clay (1995); he points out that the terms "fishway" and "fish ladder" are used in North America, whereas "fish pass" is used in Europe. Fish passes can also be classified according to the behaviour of fish, i.e. some types of fish passes enable fish to swim upstream under their own effort whereas fish lifts and fish locks lift the fish over an obstruction.
The wide variety of fish pass designs has been reviewed periodically (Orsborn, 1987; Larinier et al., 1994; Clay, 1995). An extensive review of fish passages in relation to dams is provided by Larinier (2001, this volume). As with the hauling of fish, substantial experience in design and operation of fish passes, dating back to early in the last century, has led to the development of standard design criteria (Clay, 1961). Modern fish passage philosophy includes at least five important general aspects as to the design and construction of efficient fish passes. First, speed and success of fish passage must be optimized to minimize delay, stress, and damage of fish. Second, the discharge through a fish pass has to be adapted to the needs of the target species and negotiated accordingly with the other potential or existing water resource users. The available volume of water should be used in an optimized way. Third, the fish pass should be optimized for the widest possible range of stream flows that can be expected at the time of migration of any of the target species. Fourth, funding for construction, operation, and maintenance should be allocated in such a way as to guaranty the best possible functioning of the fish pass. Lastly but not least, safe downstream migration has to be considered. Optimising the first element, i.e. effectiveness, is a challenge especially if the goal is to pass a variety of fish species that have different behaviours, sizes, and swimming abilities, but it is today technically feasible by selecting the most appropriate type of fish pass for a given situation. Unfortunately, the most successful and cost-effective fish passes are often those designed to allow passage of one or a few target species during a narrow time period, e.g. a run of anadromous fish that have a uniform size and predictable behaviour. Although some species are reluctant to pass through fish passes, in many cases species-specific structural modifications of the fish pass can encourage passage. The improved fish lift at Golfech on the Garonne River in France is a good example of how to pass a "difficult" species, i.e. Allis shad, Alosa alosa (Larinier et al., 1994). There also exist numerous examples of non-target fish species using fish passes to surmount obstacles (Schwalme and Mackay, 1985; Slatick and Basham, 1985).
Costs must not be a criteria for not choosing the optimal solution; the "User-Pays Principle" (cf FAO, 1997) must be applied, i.e. users of the water and the basin should minimize any deleterious effects and contribute to the mitigation of any impacts of their activities and to rehabilitate the systems when the need for their activity has ceased. In other words, the users of the water resource impeding fish passage by putting an obstacle should bear the expenses of carrying out the measures to ensure free and unhindered passage, both upstream and downstream, through a well functioning fish pass and/or by-pass and pay for rehabilitation, e.g. decommissioning of a dam, at the end of an activity. Furthermore, the users have to prove the effectiveness of the fish pass facility through monitoring at their costs. Where a fish pass does not prove fully effective, improvement is needed. Ideally, a fish pass should be permanently open to passage; if this is not possible, the migration periods of the different species have to be taken into consideration to ensure that the facility is functional at critical times.
In fact, fish passage facilities have been installed in many rivers where migratory species are major components of the fish assemblage. These facilities may include fish passes, fish elevators, fish locks, powerhouse collection galleries, tailrace fish diversion screens, spillways or outlet works that provide water to attract fish into fish passage facilities, and juvenile by-pass facilities. In addition to these physical facilities, other activities to facilitate passage may include transporting fish by truck and barge, increasing streamflow through reservoirs during migration to facilitate downstream movement of juveniles, modifying spillways to reduce nitrogen supersaturation during times of spill, and adjusting or reducing spill to help mitigate nitrogen supersaturation.
Trapping and hauling is a labour-intensive mitigation measure that can be used when fish need to be transported long distances upstream or around a large number of obstacles. Upstream-moving fish may be collected at a single location (e.g. the farthest downstream dam) and transported by tank truck to upstream releasing locations. The techniques and factors important to the survival of transported fish are relatively well understood based on experience with hatchery fish. However, it is more challenging a method for moving wild fish past a dam because collection is more difficult under the given hydrological circumstances below a dam, and target fish may concentrate near the dam for only short periods of time.
Dependence upon technology to provide passage around dams has not always been successful. On some occasions where upstream fish passage facilities have been provided, migration delays and mortality of adults persist. Poorly designed fish passes can inhibit movement of adults upstream, causing migration delays, increased pre-spawning mortality, and reduced reproductive success of the fish that eventually reach their spawning grounds.
The characteristics of the attraction current, and thus especially the design of the entrance, are very important to attract and guide adults into passage facilities. To attract fish, appropriate flow and water temperature must be available in the channel area adjacent to and immediately downstream from the entrance to the passage facility. These attraction currents are crucial in leading the upstream migrant fish into the fish pass from the tailwater area or open-river. Uncontrolled water spills at weir shutters or sluice gates can mislead the fish with a risk of injury. Relatively large amounts of water can be required to create suitable flows (30-60 m3 s-1 for salmonids), and these may be supplied by large capacity pumps, a hydroelectric unit whose draft tube discharges into the passage approach channels, or by direct diversion of water from the headwater to the approach channels through gravity supply systems.
The modus operandi of the turbine(s) could be a determining factor in fish finding the entrance into a fish pass. It is a well-known fact that optimal positioning of the entrance(s) of a fish pass relative to the turbine outlet is critical. Less, however, is known about the importance of sound turbine management on fish passage, e.g. the number of turbines operating at a given time, the position of operating/not operating turbines relative to the entrance (turbine next to entrance turned on or off, or running at reduced rate), the modulation of the turbines (degree of flow-through compared to maximal capacity). Further research and tests are needed to clearly determine the degree of influence of turbine management on fish pass effectiveness and efficiency. Until more is learned, it is recommended that a diversity of turbine operation cycles be provided during migration periods to benefit the greatest number and species of fish. The commanding and operation cycle of the turbines should therefore be adjusted to the needs of the migrating species. The turbine's rotation direction, and thus the orientation (spin) of the turbine outflow, relative to the entrance of the fish pass could also turn out to be an important factor. Tests of a prototype fish collection facility (a sort of collection gallery) with three entrances at one hydropower production site on River Lahn in Germany suggest that it was likely that the spin of the water coming out of the turbine determined the attractiveness of one entrance location compared to another (Adam and Schwevers, 1998). For new dam projects, the effect of the rotation direction of the turbine(s) on fish should be studied beforehand (maybe as part of the EIA).
Dams are also barriers to downstream passage of juveniles and adults that have spawned in upstream areas (e.g. salmonids and some potamodromous species) or that are on their downstream spawning migration (e.g. eels). The size, morphology, and retention time of the reservoir may limit downstream passage of juveniles through increased migration time caused by reduced currents, exposure to less favourable water quality and habitat conditions, and increased exposure to predation. At dams, injury and mortality of juveniles occurs because of passage through turbines and sluiceways. Impact with turbine blades, rough surfaces, or solid objects can cause death or injury. Changes in pressure within turbines or over spillways also can result in death or injury. Juveniles, frequently stunned and disoriented as they are expelled at the base of the dam, are particularly vulnerable to predation. Below hydroelectric facilities, nitrogen supersaturation may also negatively affect migrating fish by causing gas-bubble disease. Mortalities from gas-bubble disease increase in years of high flow and high spill. The severity and outcome of gas-bubble disease depends on the level of dissolved-gas supersaturation; duration of exposure to supersaturated water; water temperature (warmer water can hold less gas, and can therefore become supersaturated at lower pressures); health and condition of the fish; and swimming depth of fish (Marking, 1988). Eels are often injured or killed during turbine passage. Most frequently, adult salmonids end up either blocked at, or squeezed onto, the trash racks, or pass the dam in spill flows. Cumulative effects may occur where migrants encounter several dams in their downstream passage.
In view of the difficulties of passing downstream juveniles migrating past large dams, particularly those with hydroelectric facilities, a number of alternative methods have been tested in order to achieve the most satisfactory and economic solution. A variety of downstream fish passage and screening devices has been used to prevent fish from being drawn into turbine intakes (see also Clay, 1995; Odeh, 1999; Larinier, 2001, this volume). However, there is presently no single fish protection system or device which is biologically effective, practical to install and operate, and widely acceptable. Spill flows, as the simplest, can transport fish over the dam rather than through turbines or other discharges, but significant damage can occur. More sophisticated devices include highly engineered physical screening and behaviour-based guidance measures. Beginning in the 1990s, the use of Strobe lights (Martin and Sullivan, 1992), underwater electronically-generated sound (Loeffelman et al., 1991; Dunning et al., 1992), and the more economic Eicher fish screen showed potential for guiding fish more safely through, or away, from turbines (Adam et al., 1991). Use of multiple, unequal sized turbines has been found to facilitate out-migration and reduce flow fluctuations below dams (Bowman and Weisberg, 1985).
Recently, progress has been made in improving fingerling by-pass facilities at small hydroelectric plants in France (Larinier and Travade, 1999), but research and testing continue. However, as the use of fingerling by-pass facilities is still in a more or less experimental stage, and until these facilities are installed at all dams in reservoir series, it may be necessary to enhance the passage of downstream migrants by induced spill of water at critical times. Further, increasing stream flows through water releases from upstream reservoirs during critical periods can allow or enhance downstream passage of drifting larvae or migrating juveniles by providing exogenous cues that trigger the onset of migration, quickening transport, providing guide flows, and maintaining suitable water quality (Karr, 1987; Berggren and Filardo, 1993; Karr et al., 1998).
Increased spillage may be used to flush fish over a dam, or through a by-pass. These measures may be especially cost-effective when the downstream migration period of a target species is short, when migration occurs during high river flows and water would be spilled anyway, or when spill flows are needed for other reasons, (e.g. to increase dissolved oxygen concentrations or maintain minimum instream flows). Although costs of construction and labour are low for these mitigative measures, the real cost factor is the quantity of spill water that is not available for power production. Suitable discharges may be specified in the user's licence invoking the "User-Pays Principle" (cf FAO, 1997), or renegotiated if the user's licence has to be renewed. As with any fish passage device, care should be taken to ensure that mortality associated with spillway passage does not exceed turbine passage mortality (see also Larinier, this volume).
Sluiceways or by-passes are used to transport fish downstream through the dam, either alone or, more commonly, in conjunction with some other mitigative measure such as screens. If fish tend to concentrate in the upper portion of the water column, they may use orifices or overflow areas leading to ice and trash sluiceways to by-pass the turbine intakes (Taft, 1990). Designing an effective by-pass for low-head dams can be relatively easy, given proper consideration of scale; Larinier and Travade (1999) provide several examples. However, at high dams or where debris or ice in the water are abundant, fish may suffer injury or mortality in the by-pass channel or pipeline. Criteria for designing effective by-pass systems have been described (Rainey, 1985; Clay, 1995).
A simple and common means of reducing fish passage through turbines is to modify the trash racks used to prevent large debris from entering the power-plant intake. One common modification is the angled bar rack, where the trash rack is set at an acute angle to the flow direction (rather than perpendicular to flow), and individual bars may also be set at an angle to the flow. Water entering the turbine must abruptly change direction as it passes through the angled bar rack. The belief is that fish can sense and avoid this change in direction of the bulk flow and will be guided downstream along the angled rack to a by-pass. Frequently, the bars within an angled bar rack are spaced more closely than in a conventional trash rack; spacing between the bars may be reduced from typical values of 8-20 cm, to no more than 2.5-5 cm. Closely spaced bars will prevent large fish from becoming entrained in the intake flow even if the behavioural guidance aspect of the device fails.
Travelling screens are also used to prevent fish from passing through the turbines. Vertical travelling screens are commonly used at steam electric power plant intakes and rotary drum screens are often used at irrigation diversions; these designs have also been modified for hydropower intakes. The most frequently studied travelling screens for hydropower applications are the gatewell screens installed at several dams in the Columbia River basin. These screens are installed in the upper portion of the turbine intake gatewell. Because some downstream migrating salmonids are surface oriented, they encounter the screen and are forced upward into gatewells, where they pass into a flume and are routed either to a collection point (for truck or barge transportation downstream) or are discharged into the tailrace to continue their downstream migration.
A variety of other fish screens have been suggested, but some are recent developments and few have received the extensive biological testing at hydropower plants that is needed to determine their general effectiveness. Inclined plane screens, vertical punched plate screens, and cylindrical wedgewire screens have been recommended (Dorratcague, 1985). One version of an inclined plane screen (known as the passive pressure or Eicher screen) has been installed in a penstock at the Elwha Dam in Washington State, USA. In this design, downstream-migrating fish can be diverted out of the penstock and into a by-pass. Studies of the diversion and survival of salmon smolts have been encouraging (Winchell and Sullivan, 1991).
Barrier nets have been tested, but have not gained wide acceptance (Taft, 1990). Deployment and maintenance can be very labour intensive. A mesh size sufficiently small to exclude a variety of fish species and sizes will also collect water-borne debris, thereby requiring cleaning and protection from wave action. Other mitigative measures depend on fish behaviour rather than physical screens to exclude fish from turbine intakes. Behavioural barriers that have been studied include electric screens, bubble and chain curtains, chemical repellents, underwater lights, and sounds. Although the results of studies of these measures have been equivocal (Mattice, 1990), some refinements of behavioural barriers continue to be examined.
The choice of mitigative measures is dependent on the species and behaviour of fish in need of protection. If the intent of the mitigation is simply to prevent resident fish from becoming entrained in the turbine intake flow, then a physical exclusion device (e.g. angled bar rack, cylindrical wedge-wire screen, barrier net) without by-pass facilities may suffice. If there is a need to transport downstream-migrating fish below the dam, then the mitigative measure must also incorporate some means of safely conducting the fish (e.g. through by-passes, trash sluices, collection and hauling). In such cases, not only the intake exclusion device but also the subsequent downstream transport measure must be evaluated for effectiveness.
Several options are available for the restoration or maintenance of aquatic and riparian habitat. One set of practices is designed to augment existing flows that result from normal operation of the dam. These include operation of the facility to produce flushing flows, minimum flows, or turbine pulsing. Another approach to producing minimum flows is to install small turbines that operate continuously. Installation of re-regulation weirs in the river downstream can also achieve minimum flows. Also, riparian improvements are important and effective in restoring or maintaining aquatic habitat.
Flow augmentation procedures such as flow regulation, flood releases, or fluctuating flow releases all have a detrimental impact on downstream aquatic and riparian habitat, but can be managed to enhance downstream conditions. A flushing flow is a high-magnitude, short-duration release for the purpose of maintaining channel capacity and the quality of in-stream habitat by scouring the accumulation of fine-grained sediments from the streambed. Flushing flows wash away the sediments without removing the gravel (unless excessive), and prevent the encroachment of riparian vegetation. Routine maintenance generally requires a combination of practices including high flows coupled with sediment dams or channel dredging, rather than simply relying on flushing or scouring flows (Nelson et al., 1987). Minimum flows are needed to keep streambeds wetted to an acceptable depth to support fish. Because wetlands and riparian areas are linked hydrologically to adjoining streams, in-stream flows should be sufficient to maintain structure and function of wetland or riparian habitat. Flushing and scouring flows may also be necessary to clean some streambeds, flood wetlands, and to provide the proper substrate for aquatic species.
Seasonal discharge limits can be established to prevent excessive, damaging rates of flow release. Limits can also be placed on the rate of change of flow and on the stage of the river to further protect against damage to in-stream and riparian habitat. Several options exist for establishing minimum flows in the tailwaters below dams (see 2.5.3). Sluicing (releasing water through the sluice gate), turbine pulsing (releasing water through the turbines at regular intervals), and small turbines (capable of providing continuous generation of power using small flows) have been used to improve flows. Re-regulation weirs installed in the streambed below the dam to capture hydropower releases can also regulate flows (Nestler et al., 1986) to produce desired water level and velocities (and to improve dissolved oxygen, as discussed earlier).
Riparian and in-stream improvements are another strategy that can be used to restore or maintain aquatic habitat. Riparian improvements may often be more effective than flow augmentation for protection of in-stream habitat (Swales, 1989). In general, improving riparian vegetation and providing greater habitat diversity are the most effective strategies (Andrews, 1988).
The physical, chemical, and biological characteristics of reservoirs are generally intermediate between those of a river and those of a lake (Thornton et al., 1990). Operation of reservoirs strongly influences their effects on the river downstream, and can alter the ecological structure within the reservoir. Releases are perhaps the most ecologically significant aspect of reservoir operation (Stra_skraba, 1999), including: (i) quantity and rate of water releases, that also affect residence time; (ii) timing of releases; and (iii) depth from which water is released which affect stratification in the reservoir and water quality downstream and in the reservoir.
Dam release patterns may be classified either as run-of-the-river or storage-release. Run-of-the-river dams typically use little or no storage volume, and tend to operate more closely to the natural flow patterns of the river. However, in some cases flow in the river bed may be reduced or eliminated by re-routing it through pipes that carry water to turbines. These reservoirs seldom have the problems associated with hypolimnetic releases described earlier, and have limited disturbances to sediment, nutrient, and seston transport. Conversely, storage-release reservoirs withhold flow for irrigation, water supply, navigation, flood-control, or hydroelectric production, and thereby alter daily, seasonal, and annual patterns. Pulsing releases, such as those resulting from hydropeaking for energy generation, typically produce the largest environmental alterations downstream. Storage-release reservoirs change the river's sediment, nutrient, and seston transport functions, and alter water quality.
Flow variability controls all physical, chemical, and biological phenomena in a river. Impoundments, particularly those of storage-release nature, reduce the annual variability of flow, although hydropower impoundments may increase diel variability. Without established minimum flows, storage-release reservoirs may virtually stop flow for hours, days, or even weeks. Some hydropower dams have underground power stations, resulting in the desiccation of a section of the river directly downstream from the dam. Resident fishes experiencing flow alterations may be affected for great distances downstream. Flow modifications affect water quality, water depth and velocity, substrate composition, food production and transport, stimuli for migration and spawning, survival of eggs, spatial requirements and eventually fish species composition (Petts, 1984). Mitigation techniques for flow alterations include construction of re-regulating weirs (Shane, 1985), low level releases to maintain negotiated minimum flows, and manipulation of cross-sectional stream geometry. However, all these generally lead to reduced variability, whereas in nature unpredictable flows seem to produce diverse communities, as flow conditions favour different organisms.
Minimum stream flows historically were set based on some hydrologic character of the river (e.g. percentage of mean annual flow). However, low-flow recommendations are often deleterious to fish and fisheries. Improvements to establishing minimum flows include Tennant's (1976) method, wetted perimeter curves (Hauser and Bender, 1990), habitat retention models (Nehring, 1979), and a physical habitat simulation model (PHABSIM; Bovee, 1982). In general, these methods vary flow to provide suitable habitat year-round. However, many methods exist for determining minimum flow regimes (see reviews by Trihey and S. Stalnaker, 1985; Morhardt and Altouney, 1985; and Estes and Orsborn, 1986).
Flow augmentation through seasonal releases of water from storage reservoirs, can facilitate movement of fish migrating upstream or downstream, and inundation of the floodplain. Often, upstream storage reservoirs can be managed for, or their operation coordinated to provide, flow augmentation. Thus, water from reservoirs positioned high in the watershed may be released during spawning periods or periods of low flow to improve flow and flooding and improve downstream habitat conditions and fish passage. Methods for improving habitat, other than flow augmentation, are reviewed by Swales (1989) and Orth and White (1993).
The physical, chemical, and biological attributes of the downstream ecosystem are dictated by whether releases are drawn from the hypolimnion, epilimnion, or from multi-levels (Cassidy, 1989). Depth of withdrawal affects water temperature, levels of dissolved gases, nutrients, turbidity, passage of toxic or oxygen-demanding materials, and biotic assemblage and diversity. Hypolimnetic releases are relatively cold, oxygen depleted, nutrient-rich, and may have high concentrations of iron, manganese, and hydrogen sulfide (Petts, 1984). Epilimnetic releases are typically less disruptive as temperature and water quality characteristics are more suitable to the downstream biota. However, because reservoirs cool and warm more slowly than streams, normal seasonal temperature patterns may be delayed by as much as 20-50 d in some latitudes (Crisp, 1977). If the reservoir constitutes a nutrient trap, reduced nutrient discharges contribute to reduced productivity downstream. Selective withdrawal over a range of depth allows for matching releases with the dynamic conditions downstream. Nevertheless, multi-level discharges may be less desirable for hydropower producers because all available head may not be usable.
Environmental problems as a result of low dissolved oxygen (DO) may be widespread in highly impounded river systems. Although releases are aerated downstream, the length of the downstream section impacted by low DO depends upon quality of the release, turbulence, and photosynthesis and respiration. Depending on these variables, as much as about 70 km may be required before oxygen tension rises to satisfactory values (Fish, 1959), about 5 mgL-1 (Coble, 1982), although this target concentration varies depending on the fish assemblage (e.g. warmwater versus coldwater). In cascades of reservoirs, low DO in one reservoir may even affect lake and release DO in another. Thus, artificial re-aeration (Bohac et al., 1983; EPRI, 1990) through selective withdrawal, destratification, oxygen injection, turbulence-production, sluicing, or hydroturbine aeration (Bohac et al., 1982, 1983) is critical in many situations.
Manganese and iron are the two metals most often associated with releases. Both metals become soluble in water in an anoxic hypolimnion and are highly toxic to fish (Doudoroff and Katz, 1953), but rapidly oxidize when DO increases to >0.2 mgL-1. Oxidized precipitates of these metals may physically damage fish. Grizzle (1981) reported effects of manganese on fish such as destruction of gill epithelium causing respiratory difficulties and suffocation; accumulation in internal organs; lesions on gills, liver, spleen and kidney. These oxidized metals also affect consumption of water by municipal and industrial users as they stain plumbing and laundry, affect taste and odour of water, and interfere with manufacturing processes. The effects of manganese and iron are significant only at concentrations greater than 1 mgL-1 (Gordon, 1983). Practices that improve hypolimnetic DO levels in reservoirs also lower manganous and ferrous ion levels. Mitigation efforts may include changes in the release regime that allow minimal or calculated discharges of hypolimnetic waters, and rescheduling of outages that require sluicing to cooler times of the year when DO is higher.
Hydrogen sulfide may reach toxic concentrations in releases and the reservoir. It occurs in the anoxic hypolimnion and accumulates as a result of metabolic reduction of sulfates by anaerobic bacteria. Concentrations >0.002 mgL-1 have offensive odour and may cause fish kills. This level has been established in the USA as constituting a long-term hazard to most fish and aquatic life (USEPA, 1986). Practices that improve hypolimnetic DO levels in reservoirs also reduce levels of hydrogen sulfide.
Supersaturation of gases (mainly nitrogen) in releases may occur from artificial aeration of the hypolimnion, cascading spillway discharges, or through discharges of turbines operating at low generation levels. Supersaturation is aided by rapid pressure decreases and rapid temperature increases. This supersaturation of water with gases is only temporary, but return to equilibrium may often take long, and thus supersaturation can occur in tailraces. At high levels dissolved N cause fatal gas bubble disease (Bouck, 1980). A maximum saturation value of 110% total dissolved gases, at the existing atmospheric and hydrostatic pressures, is generally considered adequate to protect salmonid and other fishes (Ruane et al., 1986). Supersaturation in waters passing over spillways may be reduced or eliminated by spillway design (Smith, 1976), and in turbines by modification of air-valve systems (Ruggles and Watt, 1975), placement of perforated bulkheads in turbine bays (USACOE, 1979), or use of degassing siphons (Monk et al., 1980).
Reducing water flow changes the landscape downstream (Simons, 1979; Reiser et al., 1989). Reductions in sediment loads caused by impoundments, prompt the river downstream to try to recapture its load by eroding the downstream channel and banks. River beds are typically eroded by several meters within a decade of dam construction; this change can extend for many kilometres below a dam, depending on the slope of the terrain. River bed deepening can also lower the water table along a river, threatening vegetation and local wells in the floodplain and requiring crop irrigation in places where there was previously no need. The depletion of river bed gravels reduces habitat for many fish that spawn in the gravely river bottom.
Channel degradation processes such as bank scouring, straightening, and deepening induced by high-speed releases usually dominate below dams, with rates of erosion typically higher than those associated with unimpounded rivers (Petts, 1984). Turbidity may be due to bank erosion, sluicing of sediments from the reservoir, annual uprooting of aquatic plants, or even unusually high primary production in the reservoir. The U.S. Environmental Protection Agency recommends that light penetration should not be increased by 10% over the unaltered stream, Hynes (1970) concluded suspended solids levels >80 mgL-1 were likely to be harmful to aquatic life, and Cairns (1968) discussed more specific standards. Scouring of the tailwater persists until the channel has modified itself enough so that flow velocities fall below the threshold for transfer of streambed sediments. Degradation may be slowed by riparian vegetation, large channel cross-section, channel slope reduction, and the presence of streambed materials too large (e.g. boulders) or too cohesive (e.g. bedrock) for flow to remove.
Alteration of channel morphology and sedimentation by releases generally results in loss of habitat heterogeneity and smothering of the benthic community. Accumulation of coarse sediments on riffles, and filling of pools effectively destroy spawning, nursery, and shelter habitat of fish (Petts, 1984; Welcomme, 1985; Nelson et al., 1987). Gross substrate deposition and transport are detectable and documentable by grading with numerical substrate codes (e.g. Brusven and Rose, 1981). These methods classify substrates into size categories, and rate of change is measured through periodic classification (USEPA, 1985).
Tailwaters immediately below dams are usually autotrophic because the reservoir accumulates nutrients originating from the watershed above the dam. However, when the discharge is nutrient-rich with a relatively low level of turbidity, production of algae may be stimulated. Farther downstream, as the river becomes more heterotrophic, photosynthesis and primary productivity will be reduced to play only a limited role. Tailwater periphyton remove nutrients from the flowing water, and serve as food for zooplankton, benthos, or various fish species. Macrophytes in the tailwater are limited to littoral areas and relatively stable pools. Because tailwaters are dynamic with respect to depth and discharge, they do not provide suitable habitat for most higher plants; however, areas that are subject to frequent inundation may support bryophytes. The sediments in most tailwaters are composed of grains coarser than those in the reservoir and usually will not support rooted aquatic plants (Petts, 1984; Orth and White, 1993).
Release patterns and quality affect downstream biota in numerous ways (Gore and Petts, 1989; Welcomme, 1985; Cheslak and Carpenter, 1990). Large flow variations may adversely affect downstream productivity by impacting spawning and disrupting benthic populations. In addition, cooler releases slow chemical and biological reactions, thus reducing productivity in the affected reach. The macroinvertebrate and fish species are typical of the natural stream system, but community structure depends on reservoir operation such as duration and quantity of low-flow releases, flood releases, etc. Although lower nutrient concentrations in releases can result in lower primary production in the tailwater, the export of reservoir plankton can compensate for this reduction by supplementing the food supply for the macroinvertebrate and fish species. Conversely, nutrient-rich releases stimulate increases and even lavish development of periphyton, algae, and macrophytes. The benthic community may shift towards grazers and collectors and experience loss of diversity as organisms depending on thermal cues for spawning, hatching, and emergence will dwindle. Large diurnal flow fluctuations can have a deleterious effect on many macroinvertebrate and fish species. While the diversity of species generally decreases, those species that are able to tolerate the large flow variations can become abundant. Macroinvertebrate densities also may increase markedly during the initial downstream water surge at the start of the generation period. Macroinvertebrate transport is greatest during generation, but many of these invertebrates may originate in the reservoir and are transported into the tailwater, supplementing the food supply for the tailwater fisheries. Non-generation periods may strand some fish and macroinvertebrate species and result in their desiccation. Assessments of effects on fish communities may be quantified using methods such as the Index of Biotic Integrity (Karr et al., 1986).
Water release patterns are a central feature in the ecology of impounded systems, and when reservoirs are constructed in series, the potential for controlling timing, volume, and quality of releases is maximized. Nevertheless, the interactions that occur among the reservoirs in a series vary depending on the characteristics of individual reservoirs. Basin morphology, reservoir siting, and release features and patterns determine the interactions among the reservoir series. Factors such as bottom releases into a shallow reservoir or surface releases into a deep reservoir will drive abiotic and biotic characteristics in the series. Nevertheless, the ecological structuring and functioning of reservoir series are poorly understood, and thus their ecological impacts and benefits are hard to manage.
Major changes in water quality and phytoplankton assemblages were noted in a cascade of seven large reservoirs in the Tietê River, downstream from São Paulo, Brazil (Barbosa et al., 1999). Much of the nutrient and sediment loads were absorbed early in the cascade, particularly in the first reservoir. In this manner, the reservoirs located higher in the cascade contributed to improved DO, reduced turbidity, and overall better water quality in succeeding reservoirs. However, a proliferation in downstream eutrophication is forecasted as reservoirs high in the cascade become hypereutrophic and their ability to store nutrients diminishes.
System regulation for quantitative aspects, such as flood-control and hydropower generation, is a widely accepted and established practice, and the same principle applies to water quality concerns. Water quality maintenance and enhancements may be possible through coordinated system regulation. This applies to all facets of quality, from the readily visible quantity aspect, to traditional concerns such as water temperature and dissolved oxygen content. System regulation for water quality is of most value during low-flow periods when available water must be used with greatest efficiency to avoid degrading reservoir or river quality. Seasonal water control plans are formulated based on current and forecasted basin hydrologic, meteorologic, and water quality conditions; reservoir trophic status; water quality objectives; and knowledge of water quality characteristics of component parts of the system. Required flows and qualities are then apportioned to the individual projects, resulting in a quantitatively and qualitatively balanced system. Computer programs capable of simulating reservoir system regulation for water quality provide useful tools for deriving and evaluating water control alternatives.
An issue likely to influence the weighing of benefits and problems of reservoir development is that of cumulative impacts. Cumulative impacts can occur over a large geographical area, over a long time frame, can be the direct or indirect effect of an activity, and can occur in an additive, synergistic, or threshold manner. Cumulative impacts can be caused by dissimilar activities or projects in the same general areas, or by different activities resulting from a single project. Thus, although an individual project may have no substantial adverse effect on the fish and fisheries of a river basin, the cumulative effect of such development throughout the river basin could be quite harmful, particularly to migratory fish. Because cumulative impact assessment involves estimating the combined effects of all past, present, and reasonably foreseeable future actions, cumulative impacts have been very difficult to assess and are generally ignored.
Synergistic relations are potentially very important. Of particular concern is the potential blockage of migratory routes caused by the construction of several dams. Of equal importance are the effects of turbine-induced mortality and predation at hydroelectric sites during downstream migration. Fish that survive turbine passage may be weakened. If given time to recover, fish could continue their downstream migration. Similarly, fish passing directly through the tailwater avoid predation and proceed downstream safely. But fish that are delayed in the tailwater due to stunning during turbine passage are more subject to predation and experience higher mortality rates. For example, chinook salmon (Oncorhynchus tshawytscha) runs originating in a river in Idaho, USA, must pass eight main-stem dams on the Snake and Columbia rivers. If we assume an average mortality at each dam of 15%, then 73% of the original stock is lost through the eight passes. Ascension up these rivers to the spawning grounds is accomplished on a tight energy budget, and each fish uses its remaining body reserves to spawn. If upstream passage requires more energy to overcome repeated man-made obstacles, a fish's energy reserves may become depleted before spawning.
Cumulative impacts are difficult to assess. However, a matrix technique has been used successfully to assess basin-wide impacts. The matrix lists proposed and existing hydroelectric projects and components of the resource that can be affected. The matrix cells are assigned a number that represents the relative magnitude of the project's impact on the resource. Some impacts may be very detrimental to the resource, while others may be less damaging because they are temporary or uncommon. Thus, each impact is weighed differently and combined into a single weighed mean, which represents the potential overall impact of the project on the fisheries resource. Because development occurs on streams with different capacities to support fish populations, projects with similar design and operational modes affect the basin's fish population differently. Thus, the weighed means may be adjusted to incorporate some indication of the value of the resource potentially impacted. An example of this approach is given by (Cada and McClean, 1985).
The rationale of managing reservoirs on a basin-scale instead of on a project-by-project basis rests on the recognition that the water and land resources of a basin form a unity, and hence must be treated as such if aquatic resources are to be preserved. River basin management and planning for a basin may broadly be conceived as an attempt to identify the best possible utilisation of the available water resources given existing hydrology, land characteristics and use, dam engineering features, and concerns about the biodiversity and fisheries. Due to the multitude of water resources development options that often exist, conflicts over the utilisation of a particular source between individual schemes and the interdependency between water and land use, river basin management is indeed a complex task. The management of a reservoir within a multiple reservoir basin requires consideration of aspects about water management elsewhere in the basin. One of the major constraints to sound fisheries development in three different Asian basins was identified to be the lack of a coordinating river basin authority (Petr, 1985).
Currently, no single technological tool, model, or set of management strategies can address all components of a river basin. The ability to develop integrative management plans on a river basin scale is particularly limited for biological components. Various water resources models capable of undertaking an integrated analysis have been developed (reviewed by Lee and Dinar, 1995); unfortunately, they generally ignore biological aspects. An exception is a model derived for the Columbia River system, USA. This model uses a system of penalty functions to define the economic, social, and environmental cost of deviating from ideal operation (USACOE, 1994).
A river's estuary is a particularly productive ecosystem. They support the richest fisheries in the world and have substantial influence on marine ecosystems (Oglesby et al., 1972; Wiley, 1976). Outflows into the oceans and seas control salinities, turbidities, and nutrient levels in estuaries. Fisheries in estuaries depend on the volume and timing of nutrients and fresh water. By interrupting downstream transport of nutrients, reservoirs may drastically affect the amount and timing of organic resources available downstream. These changes can restructure biotic communities and fisheries. Declines of the pelagic fishery of the entire Eastern Mediterranean Sea have been traced to nutrient trapping by Lake Nasser reservoir (Halim et al., 1995). Reductions in nutrients due to dams on the Danube, Dnieper, Dniester, and Don rivers in Europe have been associated with reductions in fisheries in the Black Sea and Sea of Azov (Tolmazin, 1979; Volovik, 1994). A similar decline occurred in water quality and fishery production in the Caspian Sea after construction of a series of dams on the Volga River (Carre, 1978). Such effects have also been noted elsewhere in the world, including the Gulf of California and San Francisco Bay in North America, and the Oosterchelde estuary in Europe (Nienhuis and Smaal, 1994). Aside from reducing nutrient loadings, reservoirs redistribute annual patterns of discharges and temper their variability, changing the timing of nutrient availability and levels of salinity, and diluting their intensity. These changes can modify the composition of fish assemblages attuned to seasonal flow dynamics, and harm fishery yield.
Moreover, inhibition of sediment transports by dams contributes to extensive losses of wetlands. Reservoir storage in the upper Columbia and Snake rivers, USA, has altered both the seasonal pattern and the characteristics of extremes of fresh water entering the estuary. Since large-scale regulation of the flow cycle began about 1969, the variation of monthly mean flow has been reduced. Flow damping has resulted in a reduction in average sediment supply to the estuary. Between 1870 and 1990, due in part to sediment input reductions, the estuary of the Columbia River lost 8 000 ha of tidal swamps, 4 000 ha of tidal marshes, and 1 200 ha of tidal flats. This has resulted in an estimated 80% reduction in emergent vegetation production and a 15% decline in benthic algal production (NRC, 1996). Vörösmarty et al. (1997) estimate that at the global level 16% of sediments are already trapped by dams.
Mitigation of such large-scale environmental effects is daunting. Releases from dams may be modified to more closely parallel natural discharges relative to timing and nutrient and sediment loads. Extensive impoundment of river basins should accommodate reduction in estuarine fisheries and mitigate by encouraging stepped-down fishing fleets and providing relief to the ensuing social and economic impacts. Water and energy conservation guidelines and standards should be developed and implemented in the basin; savings from conservation programmes should be used to restore optimum stream flows. Options for dam removal should be pursued where feasible, to return downstream reaches to conditions more similar to environmental baseline conditions.