Hydroelectric projects potentially affect many aspects of water quality, as outlined in the Project Effects Matrix (Part I). The Clean Water Act of 1972 (33 USC § 1251 et seq), originally adopted in Public Law 92-500, requires each state to develop criteria protective of water quality. To do this, the Environmental Protection Agency and states examine the effects of specific pollutants on plankton, fish, shellfish, wildlife, plants and recreational activities and determine the levels of pollutants that can exist without harming human and aquatic life. The states then determine the most beneficial designated use(s) of a particular water body (most often aquatic species and drinking water supply), and adopt water quality criteria designed to protect these uses as well as more general provisions that protect against water quality degradation.
Dams, turbines, power lines, roads, and operations and maintenance all have some potential impact on water quality. Although watershed land uses and the potential impacts of past and ongoing discharges of anthropogenic contaminants must be taken into consideration in the development of any water quality sampling plan, the most common effects of hydroelectric projects are changes in the physical structure of the aquatic ecosystem. The most pronounced effects associated with hydroelectric operations are related to dams, such as increased sediment accumulation, alterations in temperature regimes in bypass reaches and the tailraces of powerhouses, reductions in dissolved oxygen, and increases in some pollutants such as hydrogen sulfide, nutrients, and manganese. A variety of approaches are discussed below to evaluate specific effects of hydroelectric projects on water quality, including:
These are not all of the potential effects of hydroelectric projects on water quality, and the methods to evaluate them by no means cover all the potential approaches that could be used to study the effects. Rather, this discussion focuses on those effects where a nexus to project effects is most distinct and where approaches to evaluate effects are not necessarily standardized. The following references are recommended for additional information on general water quality and limnological monitoring techniques that may be employed to assess both existing water quality problems related to hydroelectric project operations, or possibly to identify emerging trends towards future degradation.
2.1.1 Selected references
Allan, J. D. 1995. Stream ecology: structure and function of running waters. First edition Chapman and Hall. New York.
Barbour, M. T., J. Gerritsen, B. D. Snyder, and J.B. Stribling. 1999. Rapid bioassessment protocols for use in streams and wadeable rivers: periphyton, benthic macroinvertebrates, and fish. Second edition. EPA 841-B-99-002. U. S. Environmental Protection Agency, Office of Water, Washington, D. C.
Wetzel, R. G. 2001. Limnology: lake and river ecosystems. Academic Press, San Diego,California.
Virtually all biological and ecological processes are affected by water temperature. Not only does temperature directly influence chemical equilibria, but invertebrate and fish communities are also extremely sensitive to temperature. In terms of effects on biota, water temperature has direct but often subtle impacts on life history timing, habitat suitability, growth rates, rates of infection, mortality from disease and toxic chemicals, and increased exposure to both native and non-native aquatic predators better adapted to altered water temperatures.
Effects of hydroelectric project operations on the natural temperature regime of riverine sites may sometimes be related to changes in riparian shading due to tree clearing for roads, power lines,and other facilities. However, the primary effects on temperature are related to alterations in water surface area, depth, and velocity due to water diversions into or out of the stream corridor, including reservoir impoundments, and conveyance through pipelines or penstocks. Changes in the water prism along a stream reach influence the balance of heat flux into (e.g., solar radiation, air convection, ground conduction) and out of (e.g., nighttime re-radiation, evaporative cooling, and ground conduction) the reach. These effects are even more pronounced at reservoir sites, where the ratio of water surface area to reservoir volume is much smaller than that found in riverine sites, which alters the rates and balance of heat exchange with the surrounding environment. Methods to assess these effects are described below.
The following reference is recommended for additional information on instream water temperature:
Poole, G. C., and C. H. Berman. 2001. An ecological perspective on in-stream temperature:natural heat dynamics and mechanisms of human-caused thermal degradation. Environmental Management 27:787-802.
Depending on the spatial distribution of local effects (e.g., due to a point discharge, reservoir, etc.) or more widely across a watershed scale, several temperature monitoring techniques may be employed. Digital temperature data loggers (thermographs) are among the most widespread instruments in use for continuous monitoring of water temperatures in aquatic ecosystems. Thermographs are generally mounted securely in a representative location (e.g., pool, riffle, etc.) for one or more seasons, and data are regularly downloaded. A number of considerations in deploying and collecting temperature data are important, including number of times temperature is monitored, interval of collection, and location of sites (Dunham et al. 2005).
In addition to direct temperature monitoring, recent remote sensing techniques such as forward looking infrared radiometry (FLIR) allow for the collection of spatially continuous surface temperature profiles over large areas at one time (i.e. synoptic sampling). This could be used both as a diagnostic tool to compare affected areas with reference areas, locate particular stream reaches with temperature in excess of established criteria or discontinuities due to dams, cooling water discharges, hyporheic exchanges, or hot springs. This information could be used to further develop a continuous monitoring program to characterize and ameliorate any identified impacts. This method has been applied throughout the United States including at the Lake Chelan Hydroelectric Project (FERC Project No. 637) on the Chelan River, Washington; within the Yadkin Pee-Dee River Hydroelectric Project (FERC Project No. 2206) study area, North Carolina; and at the Norway and Oakdale Hydroelectric Project (FERC Project Nos. UL00-2, UL00-1) in the Tippecanoe River basin, Indiana.
2.2.1.1 Selected references
Dunham, J., G. Chandler, B. Rieman, and D. Martin. 2005. Measuring stream temperature with digital data loggers: a user’s guide. General Technical Report RMRSGTR-150WWW. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fort Collins, Colorado.
Torgersen, C. E., R. N. Faux, B. A. McIntosh, N. J. Poage, and D. J. Norton. 2001. Airborne thermal remote sensing for water temperature assessment in rivers and streams. Remote Sensing of Environment 76: 386-398.
Information gathered during water temperature monitoring can be used to model temperature effects of dams, and to develop strategies to ameliorate temperature impacts related to project operations. Stream temperature modeling can be accomplished on both reach and basin scales. The particular choice of modeling scale is most often determined by the question at hand which, for example, could range from temperature effects related to road crossings to basin-wide issues such as forest management practices.
Reach-scale models are often used to examine localized phenomena such as alterations in the riparian canopy (e.g., roads, power lines, etc.) or cold and warm water discharges. Commonly used physical models include SSTEMP (Bartholow 2002), HeatSource (Boyd and Kasper 2003), and others that were developed from channel and valley geometry, shade estimates, hydrology, and meteorology (Brown 1969). Although they may be calibrated accurately to a known temperature record, reach-based models are sometimes unable to adequately account for large changes in model inputs outside of the validated calibration ranges, or account for other effects such as tributary mixing and hyporheic exchanges.
Basin-scale models are more often used to examine land use effects on temperature regimes, but may be applied to evaluate effects of distributed project features on downstream temperatures. Common network-based models, including SNTEMP (Theurer et al. 1984), HSPF (Bicknell et al. 1997), QUAL2k (Chapra et al 2005), and others evaluate entire watersheds by calculating the temperature of individual reaches and then hydraulically routing the heat downstream. Because network-based models require substantial field-measured input data to operate and frequently do not explicitly evaluate uncertainty, more sophisticated landscape Geographic Information System (GIS)-based models such as BasinTemp (Allen et al. 2005) have been employed over larger watersheds where temperature data may be sparse.
Water temperature modeling approaches have been applied at numerous hydroelectric facilities throughout the United States including at the Cooper Lake Hydroelectric Project (FERC Project No. 2170) on Cooper Creek, Alaska; Box Canyon Hydroelectric Project (FERC Project No. 2042) on the Box Canyon Reservoir, Washington; and at the Lake Chelan Hydroelectric Project (FERC Project No. 637) on the Chelan River, Washington.
2.2.2.1 Selected references
Allen, D., W. Dietrich, P. Baker, F. Ligon, and B. Orr. 2005. Development of a mechanistically based, basin-scale stream temperature model: applications to cumulative effects modeling. Technical Report PSW-GTR-194. USDA Forest Service.
Bicknell, B. R., J. C. Imhoff, J. L. Kittle, Jr. A. S. Donigian, and R. C. Johanson. 1997. Hydrological Simulation Program, FORTRAN: User's manual for Version 11. Report EPA/600/R-97/080. U. S. Environmental Protection Agency, National Exposure Research Laboratory.
Boyd, M., and B. Kasper. 2003. Analytical methods for dynamic open channel heat and mass transfer: methodology for heat source model. Version 7.0. Watershed Sciences, Portland, Oregon.
Brown, G. W. 1969. Predicting temperatures of small streams. Water Resources Research 5: 68-75.
Chapra, S. C., G. J. Pelletier, and H. Tao. 2005. QUAL2K: a modeling framework for simulating river and stream water quality. Version 2.02: documentation and users manual. Civil and Environmental Engineering Department, Tufts University, Medford, Maryland.
Sinokrot, B. A., and H. G. Stefan. 1993. Stream temperature dynamics: measurements and modeling. Water Resources Research 29: 2299-2312.
Theurer, F. D., K. A. Voos, and W. J. Miller. 1984. Instream water temperature model. Instream Flow Information Paper No. 16, FWS/OBS-84/15. U. S. Fish and Wildlife Service, Western Energy and Land Use Team, Washington, D. C.
Additional information on water temperature modeling >>
Lakes and reservoirs have distinct structures that arise by the basin geology, morphometry, hydrology, and meteorology that affect the distribution of light, heat, and water circulation and exchange. Changes in instream water quality are inevitable when free-flowing water becomes impounded as deeper, still waters within project reservoirs, and then back to riverine environments downstream of sluice gates, spillways, and generator tailraces. If the water supply was artificially impounded as part of the original hydroelectric project development, water and power demands along with the need for flood control very often dictate the regulation and control of stream flow. Variations in flows and the large size and variations in reservoir storage volumes often disrupt natural processes that would occur if water were not impounded, or that occur in the present day upstream and downstream of the impoundment.
In general, the water column in most reservoirs has a characteristic temperature structure that is independent of the size of the basin. In summer, thermal stratification occurs, with warmer water overlying colder and denser water, much like summertime thermal stratification found in many temperate lakes (Figure 1). Depending on the flow and temperature of water entering the reservoir and the particular configuration of the reservoir inlet and outlet structures, this vertical zonation very often prevents water exchanges between warmer surface waters (epilimnion) and bottom waters (hypolimnion). Depending on the overall aquatic and upland productivity of the system, this zonation may result in hypolimnetic oxygen depletion due to the decay of natural organic matter, with subsequent increases in metals, and increased rates of nutrient exchange at the sediment water interface. Releases of this impounded water can effect downstream populations of fish and other aquatic organisms due to decreased levels of dissolved oxygen; increased levels of metals, nutrients, and turbidity; or altered temperatures.
Various approaches for evaluating reservoir temperature and chemical structure are used by researchers, the most common of which are discussed below, including lake and reservoir water sampling, sediment sampling, and bioindicator monitoring.
The following references are recommended for additional information on reservoir temperature and chemical structure:
Horne, A. J. and C. R. Goldman 1994. Limnology. McGraw-Hill. New York.
Wetzel, R. G. 2001. Limnology: lake and river ecosystems. Academic Press, San Diego, California.
Vertical water samples from lakes and reservoirs can be used to determine the characteristic structures of lakes in summer and winter and to determine important transition periods such as fall overturn, when bottom waters are rapidly exchanged with surface waters. To represent reservoir water quality and structure, in-situ profiles of physical water quality (e.g., temperature, dissolved oxygen [DO], conductivity, pH, turbidity) are collected from the reservoir surface to near the bottom of the deeper portions of each reservoir using available multiple parameter instruments (e.g., HydroLab, YSI). Depending on the resulting temperature and dissolved oxygen profiles and the current and past land uses in the basin, more intensive grab sampling and analytical monitoring of nutrients and toxic substances may also be conducted.
This method has been applied at numerous facilities including at the Niagara Power Project (FERC Project No. 2216) in Lewiston Reservoir, New York; at the Packwood Lake Hydroelectric Project (FERC Project No. 2244) in Packwood Lake, Washington; and at the Klamath Hydroelectric Project (FERC Project No. 2082) in the J. C. Boyle Reservoir, Oregon.
2.3.1.1 Selected references
Horne, A. J. and C. R. Goldman 1994. Limnology. McGraw-Hill. New York.
Wetzel, R.G., and Likens, G.E. 2000. Limnological analyses. Third edition. Springer, New York.
While hydroelectric project operation and maintenance is generally not associated with the discharge of sediment-bound contaminants, changes in the water column mixing discussed above coupled with the high sediment trapping efficiency of reservoirs can lead to increased concentrations of many toxic substances. Synthetic organic chemicals containing chlorine, including carbon tetrachloride and trichloroethylene, were available commercially by 1925, and the insecticidal properties of these compounds (e.g., dichloro-diphenyl-trichloroethane [DDT]) were recognized around 1940. Because of their persistence in the environment, the legacy of these and more recently developed compounds must be considered in any sampling program.
Depending on past and present land uses in both reservoirs and upstream areas, transported and deposited sediments act as sinks for contaminants in the reservoirs due to their high adsorption capacity, and in agricultural and industrial areas, may contain excess nutrients, bacteria, oil and grease, heavy metals, herbicides, pesticides, polychlorinated biphenyls (PCBs), and other materials.
If water quality samples identify analytical parameters above applicable water quality criteria or there is a known past or present activity in the watershed that may contribute to sediment contamination, a number of sediment sampling techniques are available to determine the extent and potential sources of the impairment. Because some toxic inorganic substances and organic contaminants associated with the sediments can also be bioconcentrated by the aquatic organisms present in reservoirs, site-specific bioindicators are often used to detect the presence of materials that may not be identified above the detection limits of most analytical water or sediment quality techniques. For example, fish tissue analysis is commonly used to detect the presence of mercury, arsenic, PCBs, and other contaminants. This method has been applied in several reservoirs in New England as part of Maine Department of Environmental Protection's Surface Water Ambient Toxic Monitoring Program; in the Clinch River and Watts Bar Reservoir, Tennessee; and at the Pueblo Reservoir, Colorado.
2.3.2.1 Selected references
Barbour, M. T., J. Gerritsen, B. D. Snyder, and J.B. Stribling. 1999. Rapid bioassessment protocols for use in streams and wadeable rivers: periphyton, benthic macroinvertebrates, and fish. Second edition. EPA 841-B-99-002. U. S. Environmental Protection Agency, Office of Water, Washington, D. C.
EPA (U. S. Environmental Protection Agency). 2000. Guidance for assessing chemical contaminant data for use in fish advisories. Volume 1: Fish sampling and analysis. Third edition. Office of Water, Washington, D. C. EPA 823-B-00-007.
Hilsenhoff, W. L. 1988. Rapid field assessment of organic pollution with a family-level biotic index. Journal of the North American Benthological Society 7: 65-68.
Slotton, D. G., J. E. Reuter, and C. R. Goldman. 1995. Mercury uptake patterns of biota in a seasonally anoxic northern California reservoir. Water, Air, and Soil Pollution, 80: 841-850.
Smith, A. G. 1991. Chlorinated hydrocarbon insecticides. Pages 731-916 in W. J. Hayes, Jr. and E. R. Laws Jr., editors. Handbook of pesticide toxicology. Academic Press, San Diego, California.
USACE (U.S. Army Corps of Engineers). 1997. Hydrologic engineering requirements for reservoirs. Washington, D. C. EM 1110-2-1420, CECW-EH-Y.
Hydroelectric projects have the potential to alter the flow of nutrients (e.g., phosphorous in sediments, algal bound nitrogen) and energy flow in river systems. Because of the long hydraulic residence times and high particle trapping efficiency of reservoirs, they often provide ideal conditions necessary for algal blooms (i.e., excess nutrients and sunlight, no flushing of algae growing in suspension) with subsequent reservoir eutrophication (water pollution caused by these excess plant nutrients) becoming a common problem associated with ongoing hydroelectric operations.
The trophic state of a lake or reservoir is based on overall system productivity and is a function of physical features (e.g., latitude and elevation as they affect sunlight and air temperatures; ratio of watershed to waterbody areas; reservoir depth; or hydraulic residence time), chemical features (e.g., nutrients, oxygen), and biological responses (e.g., primary productivity, zooplankton and fish assemblage food webs and biomass). In low-oxygen conditions of many reservoir sediments, the adsorption bond between phosphorus and sediment particles becomes unstable and often results in the transformation and release of the adsorbed phosphorus into orthophosphate (i.e. previously precipitated phosphorous becomes dissolved into the water column again). Phosphorous is often a limiting nutrient and when orthophosphate it allows an increase it algal productivity. The decomposition following algal blooms absorb oxygen out of the water, leading to an oxygen deficit. Seasonal changes in reservoir stratification are often accompanied by periods with higher and lower amounts of oxygen, leading to periods of adsorption and desorption of phosphorus, or "internal cycling," which can lead to large algal blooms during spring and fall when large exchanges of surface and bottom waters occur.
Although the most common nutrient limiting the primary production in freshwater is dissolved phosphates, nitrogen compounds can limit production in alpine and arid climates where many hydroelectric developments are situated and where watershed productivity and soil nitrogen retention is low. This has implications for both the influence of air deposition of nitrogen compounds from atmospheric sources, and also of fish passage on nitrogen arriving in river systems with migratory salmonid populations.
The following references are recommended for additional information on eutrophication:
Horne, A., and C. Goldman. 1994. Limnology. McGraw-Hill, New York.
Triska, F. J., J. R. Sedell, and S. V. Gregory. 1982. Coniferous forest streams. Pages 292-332 in R. L. Edmonds, editor. Analysis of coniferous forest ecosystems in the western United States. Hutchinson Ross, Stroudsburg, Pennsylvania.
USACE (U.S. Army Corps of Engineers). 1997. Hydrologic engineering requirements for reservoirs. Washington, D. C. EM 1110-2-1420, CECW-EH-Y.
Vollenweider, R. A., and J. Kerekes. 1982. Eutrophication of waters: Monitoring, assessment and control. OECD Cooperative programme on monitoring of inland waters (Eutrophication control),Environment Directorate, OECD, Paris.
Wetzel, R. G. 2001. Limnology: lake and river ecosystems. Third edition. Academic Press, San Diego, California.
Various approaches for evaluating nutrient exchange and reservoir eutrophication are used by researchers, a few of which are discussed below.
To assess the relative productivity of hydroelectric project reservoirs as compared with pristine sites, one commonly used tool is the trophic state index (TSI). TSI values are based on the relationship between nutrients (as measured by total phosphorus), algal biomass (chlorophyll-a), and water clarity (Secchi disk depths). A regular reservoir sampling program (e.g., quarterly, monthly) may be used to develop background data to determine the TSI. In addition to direct reporting of the TSI, comparisons to regional and national reference values may be developed to assess the severity of any identified impairment of water quality.
This method has been applied in numerous reservoirs including those within the Roanoke River Basin as part of the Roanoke River Basinwide Water Quality Management Plan, North Carolina and Virginia; at the Flaming Gorge Reservoir, Utah; and at Webbers Falls Reservoir, Oklahoma.
2.4.1.1 Selected reference
Carlson, R. E. 1977. A trophic state index for lakes. Limnology and Oceanography. 22(2): 36-369.
Stable isotope analysis has increasingly been used to determine trophic relationships in aquatic food webs. In addition to assessing natural or anthropogenic sources of nutrients associated with reservoir eutrophication, stable isotopes can be used to identify subtler food web effects related to the natural flow of nutrients from headwater reaches downstream. For example, in river systems that historically or currently support salmonid populations, the amount of nitrogen arriving in the form of post-spawning carcasses has been hypothesized as a major source of natural productivity to headwater reaches. Because dams often impede or block fish migration routes, stable nitrogen isotope ratios may be used to indicate the relative contribution of food web nitrogen in biota arriving from atmospheric, biotic, or industrial sources. This analysis has also been used to document the historical distribution of anadromous fish, as described in Section 4.0.
2.4.2.1 Selected references
Finlay, J. C., S. Khandwala, and M. E. Power. 2002. Spatial scales of carbon flow in a river food web. Ecology 83:1845-1859.
Gende, S. M., R. T. Edwards, M. F. Willson, and M. S. Wipfli. 2002. Pacific salmon in aquatic and terrestrial ecosystems. BioScience 52: 917-928.
Hobson, K. A. 1999. Tracing origins and migration of wildlife using stable isotopes: a review. Oecologia 120:314-326.
McCutchan, J., W. Lewis, C. Kendall, and C. McGrath. 2003. Variation in trophic shift for stable isotope ratios of carbon, nitrogen, and sulfur. Oikos 102: 378-390.
Post, D. M. 2002. Using stable isotopes to estimate trophic position: models, methods, and assumptions. Ecology 83: 703-718.
Vander Zanden, M. J., and J. B. Rasmussen. 1999. Primary consumer d13C and d15N and the trophic position of aquatic consumers. Ecology 80: 1395-1401.
Vander Zanden, M. J., and J. B. Rasmussen. 2001. Variation in d15N and d13C trophic fractionation: implications for aquatic food web studies. Limnology and Oceanography 46: 2061-2066.
Dissolved gas supersaturation, which is associated with powerhouse discharges due to increasing gas solubility with increasing hydrostatic pressure within hydroelectric project penstocks, received considerable attention in Canada and the United States. Elevated total dissolved gasses (TDG) can lead to a physiological condition known as gas bubble trauma in aquatic biota, which can be harmful or even fatal to aquatic organisms, as demonstrated by a number of significant fish kills in the Columbia and Snake rivers. As TDG rises from saturation to levels above 110-120 percent of saturation, there is a significant driving force of dissolved gasses into body tissues of fish and other organisms, leading to stress and mortality in some cases. Although many fish forage in or otherwise occupy deeper layers where this driving force is lower due to higher hydrostatic pressure, many resident fish use shallow waters to carry out their life cycles. Invertebrates also can develop bubbles in supersaturated water and lose the ability to swim normally, which may result in long term shifts in invertebrate community structure below powerhouse tailraces.
While modern runner and tailrace designs often can reduce the potential for TDG problems, monitoring is generally conducted in systems with hydrostatic head as little as 8 m (25 ft) because of the potential for attaining equilibrium gas concentration in excess of 110% of saturation. Reconnaissance-level surveys may be conducted using barometrically compensated dissolved gas pressure sampling devices (e.g., HydroLab and others) to determine total dissolved gas pressure. Depending on the initial levels found, the need for additional sampling or continuous monitoring during normal project operations may be assessed.
2.5.1.1 Selected references
Fidler, L. E. and S. B. Miller. 1994. British Columbia water quality guidelines for dissolved gas supersaturation. Report to BC Ministry of Environment, Canada Department of Fisheries and Oceans Environment. Aspen Applied Sciences, Ltd., Valemont, B.C.
ISAB (Independent Scientific Advisory Board). 1998. Review of the U.S. Army Corps of Engineers' Capital Construction Program, Part II, B: Dissolved Gas Abatement Program. Prepared for the Northwest Power Planning Council and the National Marine Fisheries Service. Portland, Oregon.
Weitkamp, D. E., and M. Katz. 1980. A review of dissolved gas supersaturation literature. Transactions of the American Fisheries Society 109: 659-702.