- 6.1 Introduction
- 6.2 Temperature
- 6.3 Dissolved Oxygen
- 6.4 Carbon Dioxide
- 6.5 Phosphorus and Nitrogen
- 6.6 Sediment Contaminants
- 6.7 Metals
- 6.8 Organic Contaminants
- 6.9 Longitudinal Gradients
- 6.10 Depth Gradients
- 6.11 Water-Quality Management
- 6.11.1 Monitoring Considerations
- 6.11.2 Destratification
- 6.11.3 Hypolimnetic Aeration-Oxygenation
- 6.11.4 Hypolimnetic Withdrawal
- 6.11.5 Guide Curve Management
- 6.11.6 Contaminants Management
Flowing water stored in a reservoir undergoes various physical and chemical transformations that can change the quality of fish habitat within the reservoir and the river downstream. The extent of the transformations is related to the water retention time in the reservoir, which is controlled by the reservoir’s storage capacity in relation to size of the watershed and extent of precipitation. Water in small and shallow navigation reservoirs in a large river generally undergo little or no transformation because of short retention time; conversely, water stored for many months or even years behind a deep storage reservoir in a minor tributary may undergo major transformations that can affect most life in the reservoir and in the river below the dam.
Water-quality degradation can vary widely in reservoirs. Dissolved gases and temperature often receive the most attention. Dissolved oxygen is necessary to sustain aquatic life, and temperature regulates biotic growth rates. Both temperature and dissolved gases control other physical characteristics of the water as well as chemical reactions and define the biotic, and more specifically the fish, assemblage that develops in the reservoir. Nutrient enrichment principally with phosphorus and nitrogen promotes excessive primary production, which can deplete oxygen (section 4). Contaminants including organic chemicals and trace metals are of concern because they accumulate within sediment and move through food chains and food webs (Erickson et al. 2008; Stahl et al. 2009). Total dissolved solids concentrations may be of interest for water supply and other uses. Turbidity is also a key water-quality characteristic because its effects on light transmission and water clarity define habitat characteristics (section 5). A major aspect of turbidity is total suspended solids, which are also a major transport and deposit mechanism for nutrients and contaminants in reservoirs (section 4). Water pH regulates aquatic chemistry, which can affect water use and habitat. The reservoir hypolimnion is often low in dissolved oxygen and can accumulate dissolved phosphorus, iron, manganese, and sulfide, which can produce water-quality problems within the reservoir and downstream if hypolimnetic water is upwelled or discharged. Upwelled phosphorus (internal loading) can cause summer-time algal blooms. Iron and manganese affect water color and can produce water treatment problems when water is withdrawn for municipal uses. Sulfide causes odor problems when it escapes during reaeration.
Water quality in reservoirs often can be controlled by depth-selective withdrawals. However, in many cases discharges create water-quality problems downstream in the tailwater (Miranda and Krogman 2014). Water-quality management alternatives include management techniques that can be implemented in the watershed before potential pollutants reach the reservoir (sections 2 and 8) and management techniques that may be applied within the reservoir. Other in-lake management practices include phosphorus inactivation and sediment oxidation, biomanipulation, hypolimnetic aeration, artificial circulation, and sediment removal and are discussed below.
A reservoir’s annual temperature regime is perhaps the most important water-quality attribute, capable of influencing various other water-quality characteristics. Thus, knowledge of the temperature regime is key to water-quality management. Vertical thermal variation in the reservoir produces density stratification (Elçi 2008). In the less dense epilimnion, temperature decreases slowly with depth as this upper layer is usually influenced by the depth to which light penetrates. The metalimnion separates the epilimnion from the more dense deepwater hypolimnion. In the metalimnion there is a thermocline where temperature changes rapidly with depth. Below the metalimnion, in the hypolimnion, temperatures again change slowly with depth.
The physical properties of water contribute to this temperature-induced density stratification (Wetzel 2001; Elçi 2008). Because water warmer than 39°F (4°C) is less dense, the warmer waters usually exist near the surface. But because water colder than 39°F is also less dense, some temperate reservoirs have colder waters at the surface mostly during the winter. Under both of these conditions, a reservoir may be thermally stratified. During the fall, the reservoir cools at the surface and experiences cooler inflows. At the surface the cooling produces shallow instability, and mixing can occur at progressively greater depths. This process continues through early winter. When the reservoir finally achieves, thorough cooling and mixing, a uniform temperature surface to bottom, the reservoir is said to be isothermal. This condition usually occurs in winter and continues until spring. These reservoirs experience one season of mixing each year (i.e., monomictic reservoir).
Besides water temperature, water density is also determined by dissolved substances in the water (Wetzel 2001). Increased dissolved solids increase water density. In reservoirs, dissolved solids are influenced by natural watershed characteristics and by anthropogenic activities in the watershed. In some reservoirs the density of deep waters is high enough that complete mixing does not occur (i.e., meromictic reservoir). Meromictic reservoirs are often deep (Wetzel 2001).
Many northern reservoirs stratify during winter (Rahman 1978). Their surface temperatures will be at or near freezing, and often the reservoir may be covered with ice, with slightly warmer and denser water toward the bottom. Because these reservoirs mix in the fall until they experience winter stratification and then mix again in the spring thaw until they restratify in summer, they have two seasons of mixing each year (i.e., dimictic reservoir). This condition also may occur occasionally during winter in shallow, isolated embayments of reservoirs in more southern temperate latitudes.
However, some reservoirs may experience short-term thermal stratification followed by frequent mixing (polymictic), or not stratify at all. Shallow depths, short retention times, and extensive wave-induced mixing may prevent stratification. If no stratification occurs, then the reservoir may consist exclusively of an epilimnion with possibly a weak thermal gradient, and the reservoir consists primarily of warm-water conditions.
Next to temperature, dissolved oxygen is another key indicator of reservoir water quality. The concentration of dissolved oxygen in the reservoir affects the capacity of inorganic substances to reduce other substances as well as the distribution of aerobic and anaerobic organisms. The dissolved oxygen demand may be classified into sediment oxygen demand and water column demand (Cross and Summerfelt 1987).
Sediment oxygen demand is the rate of oxygen consumption by bacteria and other organisms that metabolize organic matter in the sediment. Sediment oxygen demand is usually greatest in the uplake region of the reservoir, or in the entrance to embayments where most sediments are deposited and water is warm and shallow (section 3). Organic matter in sediment in these upper regions is high, and hypolimnetic dissolved oxygen can be depleted, particularly given that the volume of water is typically small because the uplake regions of a reservoir are often shallow (USACE 1987c). Nevertheless, these upper regions do not always stratify because their shallow depth may allow mixing by wave action and sometimes by limited flows.
Water column oxygen demand is the rate of oxygen consumption by bacteria and other organisms that metabolize organic matter in the water column above the sediment. Density flows transport oxygen-demanding material into the metalimnion and hypolimnion and entrain reduced chemicals from upstream anoxic areas to augment water column oxygen demand. Organic compounds, including those in dead organisms from the epilimnion, settle more slowly in the metalimnion and hypolimnion because of increased water density caused by lower temperatures. Because this organic matter can remain in the metalimnion and hypolimnion for a long time, decomposition occurs over a long period, exerting long-term oxygen demands (USACE 1987c).
A threshold concentration of 4–5 ppm often is used to set dissolved oxygen water-quality standards. Hypoxia results when dissolved oxygen concentrations fall to less than 2 ppm, which is generally accepted as the minimum level required to support most animal life. When dissolved oxygen levels become severely depressed or anoxic (near 0 ppm), anaerobic conditions occur. Although anaerobic conditions typically occur in near-bottom waters, it can extend upward through the entirety of the water column. Hypoxia has been shown to be an endocrine disrupter in fish, which impairs fish reproduction (Wu et al. 2003).
Inorganic carbon derived from carbon dioxide is used by plants to produce organic matter. Inorganic carbon also can control the pH and buffering capacity of aquatic systems. Inorganic carbon occurs in equilibrium in three forms (Figure 6.1): carbon dioxide, bicarbonate ions, and carbonate ions (USACE 1987c). When plants use inorganic carbon to create organic compounds through photosynthesis, the pH goes up, and the concentrations of carbon forms shift from carbon dioxide to bicarbonate ions and to carbonate ions (Wetzel 2001). The scope of this pH rise and shift in the forms of carbon indexes the buffering capacity of the water. A system with low buffering capacity (i.e., with low alkalinity) is likely to have larger fluctuations in pH and shift quicker from carbon dioxide to bicarbonate ions and carbonate ions than systems with higher alkalinity (Wetzel 2001).
Phosphorus is needed by plants and animals to build enzymes and to store energy in organic compounds, whereas nitrogen is needed to build protein. As a gas, nitrogen is important to water quality mostly when there is too much of it (i.e., super- saturated). Supersaturation can cause injury or death of aquatic organisms, including fish. This is not a common problem in most reservoirs, but supersaturation can harm fish below some hydroelectric facilities (Weitkamp and Katz 1980). Phosphorus and nitrogen are typically the key nutrients controlling primary production and are considered in detail in section 4.
Sediment plays a significant role in shaping the water quality of reservoirs. Benthic habitats are an environmental sink for many contaminants, as many pollutants sink through the water column to bond with sediment particles (Horowitz 1985; Mulligan et al. 2001). Because sediment is also an important biological habitat, uptake of toxicants into the food web is influenced by toxicant concentrations in sediment. Pollutants that settle out into sediment exist in an equilibrium state with the water above, but this equilibrium may be altered by natural and anthropogenic environmental disturbances (Theofanis et al. 2001; Jaglal 2009).
In polluted waters, pollutants mainly are found adsorbed by particles and bound to organic sediment (Chapman 1992; Baldwin et al. 2002). Changes in environmental conditions alter the various phases of pollutants found on particulates, sometimes causing the pollutants to be released into solution. Various forms of organic matter can be degraded under oxidizing conditions, releasing bound pollutants (Chapman 1992). Particulate pollutants may also become soluble within the acidic digestive tracts of detritivores such as gizzard shad, releasing these substances within the animal and making bioaccumulation through predation possible (Eagles-Smith et al. 2008).
As a result of changing environmental conditions, including dissolved oxygen and temperature, there is an internal recycling of pollutants in the sediment and hypolimnion (Baldwin et al. 2002). The processes are complex and poorly understood. The greatest amount of information is available for mercury and phosphorus. The transfer of mercury from sediment is mediated by bacteria that convert sediment-bound mercury to soluble mono-methylmercury or to volatile di-methylmercury, depending upon the pH (Erickson et al. 2008). The recycling of sediment-bound phosphorus (i.e., internal phosphorus loading) is particularly important because it may increase the rate of eutrophication within a reservoir. Many environmental and physical processes are involved in the release of phosphorus. A common process is the release of phosphorus bound to iron oxide under reducing conditions found in sediment. When bottom waters are anoxic, interstitial phosphate diffuses to the overlying water, increasing the rate of eutrophication.
Natural waters normally contain low concentrations of metals, but anthropogenic sources increase the concentrations above natural levels (Rosales-Hoz et al. 2000). Heavy metals often are discharged or leached from industrial point sources, mining operations, municipal wastewater, landfills, and the atmosphere. The problem is aggravated by the fact that there is no natural degradation process for eliminating metals from the environment. Metals shift from one compartment within the aquatic ecosystem to another, including biota, often with detrimental effects (Hart and Lake 1987). Where sufficient accumulation of metals within biota occurs through food chain transfer, there is an increasing toxicological risk through fisheries. As a result of adsorption and accumulation, the concentration of metals in sediment may be much higher than in the water above.
The behavior of metals in natural waters is influenced by the substrate sediment composition, suspended sediment composition, and water chemistry. Sediment composed of fine sand and silt will generally have higher levels of adsorbed metal than will quartz, feldspar, and detrital carbonate-rich sediment (Yu et al. 2012). Metals also have a high affinity for humic acids, organo-clays, and oxides coated with organic matter (Connell and Miller 1984). The water chemistry of the system controls the rate of adsorption and desorption of metals to and from sediment. Metals may be desorbed from the sediment if the water experiences decreases in pH, such as when anoxic conditions develop. Desorbed metals return to the water, where they recirculate and may be assimilated by the biota.
Thousands of organic compounds enter water bodies as a result of human activities. These compounds have a wide variety of properties and many may be toxic. Common organic pollutants include mineral oils, petroleum products, phenols, pesticides, polychlorinated biphenyls, and surfactants (Perelo 2010). Although some of these degrade rapidly in the environment, others accumulate in bottom sediment and bioaccumulate to toxic concentrations within the food web.
The water quality in the reservoir is related to various longitudinal, morphological, and hydrological processes. In their uplake regions, reservoirs are shallower and narrower and may maintain higher levels of flow, all of which may affect water chemistry. Moreover, these upper sections tend to trap large organic matter and debris particles that generally settle in uplake deltas with coarse inorganic particles such as sands and gravels. Finer particulate inorganic and organic matter tends to settle farther down the reservoir. Overall, the smaller the particle size, the greater the surface area to volume ratio. This increased ratio increases the sorptive capacity for transporting phosphorus, organic carbon, metals, and contaminants (USACE 1987c). Clays have a high sorptive capacity, whereas sand has essentially no sorptive capacity. Consequently, nutrients, metals, and contaminants may move into and out of the reservoir joined to fine silts and clays. Longitudinal patterns in hydrology, morphology, and settling can cause longitudinal water-quality clines.
Development of anoxic conditions within the reservoir often follows a longitudinal pattern. The pattern is dictated by local morphometry conditions and nutrient deposition patterns. Anaerobic processes may begin in the uplake portions of the reservoir if organic matter accumulations from the inflow are high and progress downstream (USACE 1987c). Conversely, anaerobic processes may develop in deep water by the dam and progress upstream (USACE 1987c). However, both upstream and downstream movement patterns can co-occur.
As dissolved oxygen concentrations decrease in the hypolimnion to about 1– 2 ppm, the oxygen conditions at the water–sediment interface can become anoxic, and anaerobic processes begin to emerge in the sediment interstitial water (USACE 1987c). Nitrate denitrification to NH4 (ammonium), N2O (nitrous oxide), and N2 occurs first (Bonin 1996; Wetzel 2001). As a result, ammonium-nitrogen can build up in the hypolimnion. Denitrification is the major mechanism for transferring nitrate out of the hypolimnion.
After denitrification, manganese compounds in the interstitial water are reduced to soluble forms able to mix with water in the hypolimnion. Thus, nitrate reduction eventually allows manganese reduction. As the system becomes further reduced, iron is transformed from ferric form to soluble ferrous forms and diffuses into the hypolimnion. As the iron is transformed, phosphorus associated with ferric compounds is released. Thus, sediment is typically a major phosphorus supplier when the hypolimnion is anoxic. During this anaerobic period, bacteria decompose organic matter into acids and alcohols, such as acetic, fulvic, humic, and citric acids and methanol (USACE 1987c).
The potential benefits of maintaining an oxic and cool hypolimnion are numerous. In some cases, cool, oxic hypolimnetic water is essential to satisfy the needs of in-lake and downstream biota. Maintenance of oxic conditions generally decreases sediment release of phosphorus and ammonia, thereby slowing eutrophication. Primary sources of phosphorus include iron complexes and microorganisms that release orthophosphate during metabolism under anoxic conditions. Oxic conditions can stimulate sediment nitrification and subsequent denitrification, resulting in a net loss of nitrogen from the system (Ahlgren et al. 1994; Rysgaard et al. 1994). Oxic conditions can also stimulate bacterial growth, resulting in increased rates of nitrogen assimilation (Graetz et al. 1973).
Water-quality enhancement opportunities occur in the watershed and in the reservoir. In the watershed, protection techniques focus on agricultural and livestock farming, forestry practices, and other human activities. Watershed management is dis- cussed in section 2.
There are in-lake enhancement techniques suitable for improving water quality. These focus on reducing the effect of an anoxic hypolimnion and management of contaminants. The technologies available to manage an anoxic hypolimnion can be separated into (1) those that prevent an anoxic hypolimnion by mixing hypolimnetic and epilimnetic waters to avoid stratification, and (2) those that prevent an anoxic hypolimnion through aeration (or oxygenation) but still maintain a distinct hypolimnion. Mixing hypolimnetic and epilimnetic water will break temperature stratification and reduce the habitat available for coldwater fishes but increase the habitat for warmwater fishes. Maintaining an oxic hypolimnion through aeration will preserve temperature stratification and provide low-temperature habitat to coolwater or coldwater fish species, prevent fish kills potentially caused by rapid turnovers that mix the epilimnion and hypolimnion, and also prevent discharge of anoxic water into the tailrace.
Hypolimnetic aeration can be beneficial to fish populations in the reservoir. During summertime in many deep reservoirs, coldwater or coolwater fish have inadequate habitat and survive between a layer of anoxic bottom water and warm surface water (Coutant 1985). By aerating the hypolimnion, fish are provided an oxygenated coolwater summer refuge. Oxic hypolimnia also may provide fish and zooplankton a dark daytime refuge in which to avoid predation (Fast 1971; Field and Prepas 1997). Also, benthos diversity and density tends to increase with oxic conditions in the sediment (Pastorok et al. 1981; Doke et al. 1995). A decrease in internal nutrient loading combined with improved zooplankton habitat under aerated conditions may cause a decrease in algal biomass or a shift to more desirable phytoplankton species (section 4).
Monitoring a reservoir’s water quality can be an expensive and time-consuming task (Bartram and Balance 1996; Green et al. 2015). Whenever possible, water quality data may be obtained by partnering with agencies whose mission is water-quality monitoring. These may include the agency that controls the water stored by the reservoir, local and state departments of the environment, federal agencies including the U.S. Environmental Protection Agency (USEPA) and the U.S. Geological Survey, and universities. Section 305(b) of the Clean Water Act requires states to develop an inventory of the water quality of all water bodies in the state and to submit an updated report to the USEPA every 2 years. This process was established as a means for the USEPA and the U.S. Congress to determine the status of the nation’s waters. The 305(b) report includes an analysis of the extent to which water bodies comply with the “fishable/swimmable” goal of the Clean Water Act; an analysis of the extent to which the elimination of the discharge of pollutants and a level of water quality achieving the fishable/swimmable goal have been or will be attained, with recommendations of additional actions necessary to achieve this goal; an estimate of (1) the environmental effects, (2) the economic and social costs, (3) the economic and social benefits, and (4) the estimated date of such achievement; and lastly, a description of the nature and extent of nonpoint sources of pollutants and recommendations of programs needed to control them—including an estimate of the costs of implementing such programs. However, the Clean Water Act is limited to waters with a significant nexus to navigable waters, and agriculture nonpoint discharges are generally exempted from regulatory oversight through the Clean Water Act.
In the absence of existing data or collaboration opportunities, agencies tasked with managing fish habitat may opt for narrowly focused monitoring programs, such as a program focused on temperature and oxygen. Suitable temperature and oxygen conditions generally will limit problems associated with toxic conditions and associated compounds, including ammonia, sulfide, manganese, and metal compounds. Monitoring of nutrients and water clarity is considered in sections 4.4.1 and 5.8.1.
Monitoring temperature and dissolved oxygen is probably most effective in May through September as reservoirs begin to warm and potentially stratify. Sample stations may be established at the deepest part of the reservoir (usually near the dam) or at the midpoint of the reservoir. Additional sample locations may be needed if the reservoir is long, there is interest in the status of various embayments in the reservoir, major inflows occur within the reservoir at various locations, or the lake is divided into significant subunits by causeways (Green et al. 2015).
Water column profiles can be taken with multiparameter sondes or other field meters (Green et al. 2015). Variables including dissolved oxygen concentration, percent oxygen saturation, and temperature can be recorded at regular depth intervals (OEPA 2010; Green et al. 2015). Other useful data often available from sondes may be recorded. The first reading may be taken at the surface (1 ft) and subsequent readings at suitable intervals proportional to depth of the sampling site.
Monitoring may identify the need to avert stratification. Mixing of the reservoir to destratify layers or prevent stratification is accomplished with three general procedures: aeration, pumping, and hypolimnetic withdrawals. Mixing will help produce nearly even oxygen and temperature conditions throughout all depths. These conditions will affect in-lake water quality and that of water released through the dam. Mixing can limit phytoplankton blooms by reducing the amount of sunlight reaching phytoplankton by causing plankton to recirculate below the photic zone. Destratification will change the amount of habitat available for warmwater and coldwater fishes. Destratification will also warm the release waters, which may affect the use downstream. Lorenzen and Fast (1977) outline some additional benefits and consequences for reservoir destratification.
Johnson (1984) and Singleton and Little (2006) provide overviews of the various types of aeration–destratification systems, how they operate, and guidelines for system selection based upon the reservoir characteristics and the goals of the system to be installed. Whereas equipment technology continues to improve, the general approaches have remained the same. In general, the system design depends on the volume of water to be mixed and the temperature and oxygen profile. Equipment alternatives are available regardless of project scale (e.g., area, depth, oxygen demand) requirements (Johnson 1984).
Destratification of impounded waters may be achieved with diffused air aeration. The diffusers are normally flexible tubes that are installed on the bottom of the reservoir (Figure 6.2). The air bubbles move upward, creating an upwelling of cold water that spreads out laterally upon reaching the surface while carrying anoxic water upward. Once cold, aerated water reaches the upper layers, it sinks back again, bringing oxygen to the hypolimnion. This action causes the reservoir to destratify. Normally, compressed air is used. In general, linear or circular diffusers strategically positioned on the reservoir bottom are supplied by a compressor located on shore or fixed within the reservoir in a floating system (Single- ton and Little 2006).
Mechanical flow pumps provide enough mixing in a local area to reduce or eliminate thermal stratification (Mueller et al. 2002; Gafsi et al. 2009). Three general types of mechanical flow pumps have been used (Gafsi et al. 2009). The first one employs a water pump located on a floating platform or on shore. A pipe extends into the hypolimnion. Water is drawn from the hypolimnion, passes through the pump, and is discharged into the epilimnion or back into the hypolimnion (Figure 6.3) where it mixes (Hooper et al. 1953; Fast 1994). The second type of mechanical water pump consists of a motor located on a moveable float. A tube extends from the float into the hypolimnion, and a propeller and shaft extend into the tube (Symons et al. 1967). The propeller draws water into the bottom of the tube, and it is forced to the surface and discharged. Symons et al. (1967) compared the efficiency of this system with diffused aeration and found the latter more efficient. A third type of pump jets surface water down into the hypolimnion to achieve a locally uniform vertical temperature profile. A limitation is that the jet may strike the bottom and cause resuspension of sediment and erosion.
Because of the energy consumption and high costs associated with running electrical aerators or pumps, several commercial companies have developed solar-powered or wind-powered aerators for use in reservoir aeration or oxygenation. The advent of these technologies has reduced the cost of treatments by eliminating the need for an electrical grid and for power—and without greenhouse gas emissions.
Solar units operate by capturing solar energy via an array of solar panels, converting the solar energy to electrical energy, and then delivering the electricity to a small motor to drive either an aerator or a pump (Figure 6.4). One of the companies currently supplying solar technology produces the SolarBee® mixer (Medora Corporation, Dickinson, North Dakota). The mixer is designed for simple handling and deployment with a small crew. Maintenance of the mixer is minimal and typically consists of removing weeds and debris from the impeller and cleaning the solar panels as needed. SolarBee® mixers have been installed to enhance water quality in many reservoirs (Figure 6.5). For example, several units were installed in Jordan Lake, North Carolina, to assist with improving lake water quality in the vicinity of water supply intakes. The aeration systems were expected to reduce dissolved manganese and iron concentrations, the proliferation of cyanobacteria blooms in the vicinity of the intake and associated taste and odor issues, and the overall reservoir water quality in the area in which the intakes are located. Follow-up monitoring suggested the mixers were only partially successful.
Windmills resemble smaller replicas of the stately old windmills that, at one time, were common across rural landscapes. The windmill blades harness winds to power a crankshaft that operates a diaphragm that pumps air or oxygen into the reservoir hypolimnion. Windmill aerators have been installed in various reservoirs operated by the U.S. Bureau of Land Management. They are more difficult to install than solar aerators, making them less mobile and requiring careful planning before installation. This technology has been extensively applied to pond management (Koenders Windmills Inc., Saskatchewan) but is just recently being applied to reservoirs and intensive evaluations are not available.
Hypolimnetic aeration–oxygenation is achievable through a variety of approaches, ranging from pumping the hypolimnetic water to the surface for aeration and returning it to the hypolimnion, to fine-pore pneumatic diffusers placed in the hypolimnion for the introduction of oxygen (Figure 6.6). These systems introduce oxygen into the hypolimnion as air or as pure oxygen. Pneumatic diffuser systems are designed such that the rising bubble plume does not mix hypolimnetic water into the epilimnion and thus prompt destratification (Singleton and Little 2006). Systems that transfer hypolimnetic water to the surface for reaeration are designed to minimize the speed at which the water is transferred so that mixing does not induce unwanted destratification (WOTS 2004).
There are a number of potential problems associated with aeration. The oxygen transfer efficiencies of most hypolimnetic aeration techniques are low, ranging from about 10% (Smith et al. 1975) to 50% (Bernhardt 1967). Thus, aeration units may need to operate at high recirculation rates, which could produce turbulence within the hypolimnion and thereby increase sediment oxygen demand (Smith et al. 1975; Singleton and Little 2006). Large reservoirs may require the installation of numerous aeration units, which potentially could produce enough turbulence to cause destratification (Heinzmann and Chorus 1994). Moreover, the introduction of compressed air that predominantly consists of atmospheric nitrogen may elevate levels of dissolved nitrogen gas in the hypolimnion and the formation of gas bubble disease in fish (Beutel and Horne 1999).
The primary advantage of hypolimnetic oxygenation over aeration is that the solubility of pure oxygen in water is roughly five times that achievable via aeration because air is about 20% oxygen. A second advantage of hypolimnetic oxygenation systems is their high transfer efficiencies (percentage uptake of delivered oxygen), which generally range 60%–80% (Speece 1994; Mobley and Brock 1995). As a result of higher oxygen solubility and higher system transfer efficiencies, size of the mechanical devices and recirculation rates needed to deliver an equivalent amount of oxygen using pure oxygen instead of air are greatly reduced. This scale reduction avoids a number of the disadvantages associated with traditional aeration systems (Singleton and Little 2006). Lower recirculation rates minimize turbulence introduced into the hypolimnion, thereby minimizing induced oxygen demand (Moore et al. 1996). High oxygen delivery rates and low induced oxygen demand allow for the maintenance of suitable levels of dissolved oxygen in oxygenated hypolimnia (Thomas et al. 1994; Prepas and Burke 1997). Smaller oxygenation systems also may be able to oxygenate large bodies of water with a reduced risk of accidental destratification. Additional advantages of hypolimnetic oxygenation include avoidance of hypolimnetic dissolved nitrogen supersaturation (Fast et al. 1975) and substantial decreases in energy use (Speece 1994).
Three general types of systems can be used: bubble plume oxygenation (Singleton et al. 2007), deepwater oxygen injection through linear or circular diffusers (Mobley and Brock 1995; Prepas and Burke 1997), and submerged down-flow bubble contact chambers (Speece 1994).
Bubble plume oxygenation works by injecting pure oxygen through a dense group of porous diffusers at the bottom of the reservoir, creating a gas–water mixture that rises and gains momentum due to a positive buoyancy flux (Singleton et al. 2007). Oxygen bubbles dissolve into a surrounding plume of rising water. The oxygenated plume then detrains and spreads out horizontally below the thermocline. Bubble plumes are generally linear or circular and inject oxygen at a relatively low gas-flow rate (Schladow 1993). These systems are most suitable for deep reservoirs where the bulk of the bubbles dissolve in the hypolimnion and the momentum generated by the plume is low enough to prevent significant erosion of the thermocline. Bubble plume oxygenation may have trouble maintaining a well-oxygenated sediment–water interface because most of the oxygen rises to the upper levels of the hypolimnion.
Linear or circular diffuser oxygenation systems consist of an extensive network of linear diffusers that release fine oxygen bubbles that rapidly dissolve into the overlaying water column (Singleton and Little 2006). The diffuser system has a few advantages over other systems. In contrast to contact chambers (section 188.8.131.52) the system does not require the pumping of water. In addition, unlike the bubble plume oxygenation system (section 184.108.40.206), at low gas-flow rates the system does not induce a large-scale vertical current of water. Thus, dissolved oxygen tends to stay deeper in the reservoir. A system installed at Douglas Reservoir, Tennessee, successfully oxygenated the hypolimnion and improved water quality of turbine discharges (Mobley and Brock 1995).
The submerged contact chamber oxygenation systems consist of a submerged cone-shaped contact chamber installed on the bottom of the reservoir. A submersible pump draws water from the hypolimnion into the top of the cone. Oxygen supplied from an onshore facility is injected at the top of the cone. The oxygenated water is discharged through a horizontal diffuser pipe. Speece et al. (1971) observed oxygen transfer efficiency in the range of 80%–90% in an experimental chamber. With the proper horizontal dispersion of reoxygenated water, a submerged chamber system can overcome potential limitations of a bubble plume or a diffuser system. These limitations include accidental destratification caused by oxygen bubbles rising through the thermocline (Speece 1994) and localized anoxia as a result of limited oxygen dispersion within the hypolimnion (Fast and Lorenzen 1976). In addition, in contrast to bubble plumes and line diffusers, horizontal dispersion sends reoxygenated water out over the sediment, thereby keeping highly oxygenated water in direct contact with the sediment and promoting a well-oxygenated sediment-water interface.
Submerged contact chamber systems have been operated successfully in various water bodies (Figure 6.7). Camanche Reservoir is a large, multipurpose reservoir in the foothills of the Sierra Nevada Mountains in Northern California. A fish hatchery just downstream of the reservoir experienced large fish kills due to hypoxic withdrawals from the reservoir. After a contact chamber oxygenation system was installed, no fish kills occurred. Spatial monitoring of dissolved oxygen showed that a well-oxygenated plume of deep water migrated about 2 mi longitudinally up the reservoir 40 days after oxygenation (Speece 1994). In Newman Lake, Washington, low oxygen levels in bottom waters during the summer resulted in a severely degraded coldwater fishery. A contact chamber oxygenation system dramatically improved bottom water quality for fish during the summer by maintaining a well-oxygenated hypolimnion (Doke et al. 1995; Moore et al. 1996).
Hypolimnetic withdrawal (Figure 6.8) is a form of selective withdrawal through the dam but with the release of water from only the hypolimnion (Nürmberg 1987; Dunalska 2001; Hueftle and Stevens 2001). Epilimnetic water, which maintains adequate concentrations of dissolved oxygen, is retained in the reservoir. The major objective is the reduction of anoxic conditions in the hypolimnion which, in turn, will limit the release of phosphorus from the sediment and reduce the cycling of nutrients to the epilimnion. On an annual basis, the volume of water released remains unchanged, but the thermal stability may be reduced by withdrawing water from the hypolimnion. The effectiveness of this approach would depend upon the reservoir’s morphology and inflow water-quality characteristics. Nürmberg (1987) evaluated hypolimnetic withdrawals in nearly 50 lakes and reservoirs and reported that withdrawals decreased summer average epilimnetic phosphorus and chlorophyll concentrations, increased Secchi disk transparency, and decreased hypolimnetic phosphorus concentration and anoxia.
Hypolimnetic withdrawal does not involve mixing the epilimnion with the hypolimnion, which is possible with aeration methods. Therefore, hypolimnetic withdrawal helps control eutrophication of the reservoir. Nevertheless, hypolimnetic withdrawals release nutrient-rich water of poor quality (i.e., low dissolved oxygen, low temperature, high dissolved solids) downstream and potentially increase eutrophication of downstream water bodies.
Many reservoirs operate under some type of water-level management plan, or guide curve, that directs seasonal change in water storage. Modification of the reservoir’s guide curve may enhance water quality. By modifying the annual distribution of retention time, undesirable water quality may be avoided or flushed out. Similarly, by modifying annual storage distribution, the intensity of stratification may be controlled. Guide curve modification for water-quality purposes is not common because such action usually imposes on other objectives of the reservoir. However, reservoir guide curves have been adjusted for other purposes, such as ensuring minimum downstream flows during low-flow periods, and adjustment for water quality is identified as an option by U.S. Army Corps of Engineers (USACE 1995). See section 7 for more on guide curve management.
Modifying the guide curve may allow the flexibility to route inflows through the reservoir to release low-quality inputs or to flush low-water-quality conditions developed within the reservoir. Low-quality inputs may include highly turbid water during the wet season. This water can be moved through the reservoir more quickly by maintaining a low reservoir volume. Undesirable water-quality conditions may develop in late summer and fall when the reservoir stratifies. Increasing flow during this period may reduce the intensity of the stratification process, although extra flows during this period may be hard to get unless water is available from reservoirs upstream.
It is also possible to operate multiple reservoirs within a river basin to meet water-quality targets at downstream points within the basin. For example, releases of good quality from one reservoir can be blended with poor-quality releases from another reservoir in an adjacent feeder stream to neutralize the negative water-quality attributes. The USACE has used multi-reservoir releases to neutralize releases with low pH and to dilute highly turbid releases. The HEC-5Q model (USACE 1986) contains algorithms to calculate release requirements from multiple reservoirs to satisfy a downstream water-quality target, but other models are available (reviewed by Labadie 2004).
Management practices applicable to improving water quality may sometimes also be applicable to management of contaminants (WOTS 2004). Heavy metals that are mobilized when dissolved oxygen concentrations are low may be managed with aeration practices. Mobility may also be controlled by adding chemical materials (amendments) to the water or sediment. A “do-nothing” alternative that allows contaminated sediment to be buried with time may be feasible if contaminants remain bound to the sediment. Other alternatives may involve water management practices such as employing suitable water residence times to allow dilution or water drawdowns to allow aeration and drying. For contaminants in sediment that are primarily cycled by biotic organisms, remediation may be possible through separation of the biota from sediment by means of capping, dredging, or isolation (WOTS 2004; Jaglal 2009).
Amendments are add-ons, usually possessing high cation exchange capacity, which can lower mobility and bioavailability of contaminants in sediment, thereby decreasing their solubility. In situ immobilization using inexpensive amendments such as minerals (e.g., apatite, lime, zeolites, beringite) is considered promising (Peng et al. 2009). Compared with the amendments used in terrestrial soils, those used in submerged sediment usually have higher sorption capacity, lower water solubility, higher stability under reducing and oxidizing conditions, and lower cost (Raicevic et al. 2006).
Decreasing the direct contact area between water and the contaminated sediment can lower the release of contaminants. Therefore, capping the contaminated sediment with sandy materials, such as clean sediment, sand, or gravel, could be an effective remediation technique (Peng et al. 2009). If properly designed, the placement of a relatively coarse-grained cap does not disturb or mix with an underlying soft, fine-grained sediment. Theofanis et al. (2001) indicated that a good cap may be about 20-in thick and that capping the sediment with sandy materials can reduce heavy metal concentration in water by about 80%. Compared with other in situ remediation methods, the capping approach is low cost. To further immobilize contaminants and enhance the cap quality, amendments (section 220.127.116.11) may be mixed into the sand cap. Jacobs and Förstner (1999) reported that fixation capacity for heavy metals and organic contaminants increased sharply after adding zeolite into a sand cap.
Phytoremediation is the use of plants to extract, sequester, or detoxify pollutants. This technology is widely viewed as an ecologically responsible alternative to environmentally destructive chemical remediation methods (Meagher 2000). Phytoremediation comprises two tiers, one by plants themselves and the other by root-colonizing microbes that degrade the toxic compounds further to nontoxic metabolites. This technology popularly is applied in terrestrial soil remediation and also shows some potential value for remediation in shallow rivers, lakes, and wetlands. At present, this technology reportedly presents good immobilization effects for zinc, iron, manganese, and cadmium in sediment (Peng et al. 2009).
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