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102 4.1.2 Marine Site Evaluation The impact on the ocean resulting from discharged waste depends on the composition and volume of the waste and on the dispersal and transport characteristics of the site selected for disposal. Clearly, the distribution, fate, and effects of waste inputs are governed by the physical, chemical, and biological processes that alter the chemical forms of the waste and their bioavailability. These processes are discussed in detail in Chapter 2. Sewage sludge has been routinely discharged or dumped in the sea at several sites along both the east and west coasts of the United States. From these disposal activities, data bases are available that can be used to begin to evaluate the biological impact of waste disposal in the marine environment. Although our discussion is focused on the impact of sewage sludge disposal, the input of chemical contaminants from the disposal of dredged material and industrial wastes will result in similar effects in the marine environment. 4.1.2.1 Nearshore Disposal Sewage sludge is discharged to U.S. coastal waters either by pipeline (commonly used on the west coast) or barge (primarily used on the east coast). Although the discharge method will generally determine the initial dilution of wastes, subsequent dispersal and transport will depend on advective processes. Along the coast of southern California, the city of Los Angeles discharges sewage sludge (1 percent solids) through a pipe 75 cm in diameter at a depth of 100 m along the rim of a submarine canyon 10 km from shore (Bascom, 1982). Initial dilution of the wastes by 102 occurs, and further dilution and transport of the waste plume are achieved by passive advection and lateral spreading (Brooks et al., 1982). Because of the highly stratified water column at this site, the waste plume usually remains at a depth below the pyanocline. Contaminants of biological concern, such as pathogenic microorganisms, trace metals, and xenobiotic organic compounds, are primarily associated with particulate material, and transport of the sludge particulates is controlled by the same phenomena as is the transport of natural sediments. Only 10 percent of the sludge solids discharged settle on the bottom within 5 km of the

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105 dissolved oxygen and accumulation of nutrients in the benthos. Boesch (1982) concluded that the altered benthic community within Christiaensen Basin is better able to cope with organic enrichment than is the indigenous community, but it is less suitable for support of higher trophic levels. Similar changes have been observed in the southern California Bight, and only generalist feeders among demersal predator populations appear to be unaffected by alterations in benthic communities (Allen, 1975; Word, 1979). In addition to the high level of concern about toxic chemicals and pathogens in the marine environment, there is also concern about the release of degradable and nondegradable organic matter and nutrients to the ocean. If these substances are discharged in sufficiently high concentrations to oceanic areas of poor dispersion and mixing energy, depletion of oxygen as a result of the high rate of microbial degradation may occur. Eutrophica- tion of coastal areas from nutrient enrichment may result in changes in species composition and dynamics of marine communities. Mearns et al. (1982) have recently reviewed the effects of nutrient and organic enrichment on marine ecosystems, focusing primarily on the data base available for the New York Bight Apex. Coastal waters of the New York Bight and adjacent estuaries receive high annual inputs of organic carbon, nitrogen, and phosphorus from multiple sources, including barge dumping of sewage sludge and dredged materials, inputs from the Hudson- Raritan estuary, and other coastal and atmospheric inputs (Mueller et al., 1976). In the New York Bight Apex, seasonal and annual variations in productivity and stratification of the water column may lead to periods of low dissolved oxygen or anoxia in the benthos, such as that experienced during the summer of 1976 following a bloom of Ceratium tripes. Nutrient enrichment has also been observed in the Southern California Bight (G. Jackson, Scripps Institution of Oceanography, La Jolla, California, personal communication) On both the southern California coast and in the New York Bight, there are many sources of contaminants and nutrients, so biological effects cannot be attributed to the impact of dumping of sludge alone. Understanding the impact of other point sources is necessary to predict overall degradation of a receiving area. Clearly, the impact of waste discharges depends on the volume and composition of waste to be discharged and on dispersal characteristics at the site of discharge. Low-volume

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106 inputs from a small coastal city will not have the same impact as inputs from a large metropolitan area, in either total volume or contaminant loading. These factors must be taken into account in future permit decisions for ocean dumping. 4.1.2.2 Deep-Water Disposal Deep-water disposal of wastes, such as sewage sludge, offers the advantages of greater dilution and dispersion, reducing the potential return of wastes to humans and reducing the potential impact on living resources in nearshore coastal areas. Two deepwater sites have been proposed for receiving sewage sludge: the 106-Mile Ocean Waste Disposal Site (Dumpsite 106) located 106 nautical miles southeast of New York Harbor on the continental slope in the northwest Atlantic at a water depth of 2,000 m; and the proposed Orange County deep-water disposal site located off the coast of southern California at a depth of 300 to 400 m. Dumpsite 106 is typical of slope water regions of the northwest Atlantic, and experience with industrial waste dumping at this site provides a background of mixing and dispersal characteristics of waste inputs. Initial dilution and dispersion of wastes will be similar to those measured for barged wastes in the New York Bight, but the greater depth and proximity to the Gulf Stream ensures greater horizontal transport (O'Connor and Park, 1982). Despite a limited number of investigations that suggest that deposition rates of sewage sludge to deep benthic areas would be minimal, no accurate information exists on the potential deposition rates of sewage sludge to deep sea benthic areas in the vicinity of Dumpsite 106, the extent of the area of deposition, or the resulting impact on benthic systems. Predictive transport models of the pipeline discharge from the proposed Orange County deep-water disposal plan indicate that initial dilution of wastes will be 5 x 102, or greater, depending on the prevailing current regime and the height to which the submerged plume may rise. Further mixing is accomplished by advection and lateral spreading. The most critical difference between other outfalls off the southern California coast and this proposed outfall is its proximity to the oxygen minimum layer, and the potential effects of high biodegradation rates on biota acclimated to a low ambient oxygen

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107 concentration (1 mg/L). Brooks et al. (1982), in conjunction with a baseline study conducted by the Southern California Coastal Water Research Project, have developed a comprehensive research program to address the feasibility and impact of this particular disposal option. The paucity of documentation for the disposal of wastes at deep-water sites makes predictions of ultimate biological and/or ecosystem effects difficult to assess; further research is required for evaluation. 4.1.2 Terrestrial Site Evaluation The use of terrestrial and affiliated freshwater ecosystems for the disposal of anthropogenic wastes involves issues that are quite distinct from those of the marine situation. Because of the more intimate contact that humans may have with the wastes, compared, for instance, with contact from open-ocean disposal, the central goal of disposal for terrestrial systems is containment of the various components of the wastes. general, a properly sited terrestrial waste disposal system incorporates an area within which impacts on the natural ecosystem are not considered important and In concern instead is focused on export to other ecological and human systems. As discussed in detail in Chapter 3, such concerns include (1) long-term environmental effects, including contamination of surface or groundwater re- sources, potential threats to human health, and secondary effects on valuable natural and agricultural ecosystems and (2) long-term commitment of land resources. Land spreading and reclamation may also be used as part of ecosystem management practices, such as the use of wastes for nutrient enrichment of park lands to enhance diversity and productivity. There are a large number of terrestrial waste disposal options currently in use (Loehr et al., 1979), such as sludge applications to agricultural land (Council for Agricultural Science and Technology, 1976), sewage treatment by means of cypress domes, other wetlands and silvicultural areas (Ewel et al., 1982), and disposal in landfill and mine reclamation areas (Sopper and Kerr, 1981). Many of the constituents of waste enter surface- water and groundwater systems (Loehr et al., 1979? either through deliberate disposal (e.g., into some rivers and lakes) or secondarily (e.g., from leachate from agri

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108 cultural and forest systems). The purpose of this section is not to discuss specific disposal methods or recipient systems but to highlight those aspects of land disposal that need to be addressed. The exports from disposal systems can be categorized as nutrients, organics, heavy metals, and pathogens. The pathways of concern for these include both those linked to other natural systems and those linked to humans. Disposal systems should be designed to prevent direct or indirect contamination of freshwater, groundwater, and estuarine systems, because such systems are characterized by extensive contact with humans, lack of containment, and high concern for system alterations (Ewel et al., 1982). With respect to nutrients, direct enrichment of the disposal area may result in positive benefits, such as increased harvest of food or wood products. This is a key aspect of the application of wastes to managed terrestrial systems in that the resource value (i.e., nutrients) of human activities is recycled. Concern develops with the inadvertent nutrient enrichment of water systems from surface runoff and via percolation of leachate to groundwater. Nitrogen and phosphorus are of primary concern, particularly the movement into ground- water of nitrates and the movement of nitrogen, phos- phorus, and oxygen-demanding organics into surface-water systems (Loehr et al., 1979). The latter is more problematical where climate, topography, and system management practices (e.g., agricultural cultivation) result in significant fluxes of runoff into streams and lakes. Movement of phosphorus into groundwater has been found to be far less significant (National Research Council, 1978~. For various terrestrial disposal systems, transport of toxic organics and heavy metals into both surface- and near-surface water systems remains an issue of concern, with respect to impacts on other natural systems and particularly with respect to pathways to humans. It is beyond the scope of this section to treat this topic in detail; rather, we will simply indicate that the waste stream must be characterized with respect to these toxicants and that their physicochemical characteristics and those of the environment are critical to determining the fate, transport, and effects of the toxicants. Transport of pathogens to humans must be addressed for any terrestrial system. Potential pathways include direct consumption of food products grown in waste-amended

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110 population- and community-level effects occurring within the area under primary impact rather than on processes, because of the limited spatial extent of the population perturbation relative to the spatial scale of most ecosystem processes. When considering the effects of waste disposal at population and community levels, the first concern is potential elimination of a species, either through direct toxicological impacts or indirect effects, such as loss of habitat or reduction in some essential resource base. The issue of the spatial extent of the impacted area versus the spatial range of the species is critical. Further, for many species the area of concern includes the range for early life stages (e.g., nesting or spawning areas), which is smaller than the total geographical range for the mature stages. Similarly, the loss of a particular habitat that is both limited in its general occurrence and spatially of the same scale as the waste-impacted area presents a problem that must be addressed, as does the elimination of unique biotic communities. Of less dramatic concern is the alteration of com- munity structure. Community structure continually changes, even without significant anthropogenic pertur- bations; thus, alteration in the community structure per se may not represent a major problem. The situation becomes important, however, if such community alterations are major in spatial extent or unidirectional change and in the relationship and distribution of the constituent species. To evaluate these alterations, one must look at the interrelationships among the species. There may be some critical species whose presence is required for a significant part of the overall community to exist. Loss of the critical species from a location will result in concomitant indirect losses or population explosion of other species. Similarly, there are critical groups of species, i.e., a number of species may function redundantly within the system. The loss of all of them would result in the loss of their critical function in the overall ecosystem. Another class of indirect effect that must be con- sidered is the impact on some species that have par- ticular aesthetic or economic value. For example, pollution-induced reduction in benthic habitat could result in depletion of fisheries, even though the fish were not directly affected by the pollution.

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111 A final issue that needs to be addressed is the recoverability of the systems under impact. Rates of recovery of defaunated or defoliated areas depend on the rates at which the area becomes habitable again and on the rates of recolonization by the biota. These values depend to a large degree on the spatial scale of the impacted area. The recovery of a small area surrounded by undamaged systems is more rapid than is recovery of large impacted areas that have relatively fewer sources of colonizers. In summary, the aspects of ecosystems that must be considered when evaluating waste impacts range from direct effects on individual species and effects resulting from interspecific interactions to effects on community structure and concomitant functional relationships. These aspects are overlaid by consideration of spatial and temporal scales. These are the types of information and understanding required, but it is quite another matter actually to attain them. For instance, while direct impacts on heavily affected areas may be readily discernible, effects of longer-term, more widely disseminated lower-level stresses are generally more difficult to detect. Such chronic stresses can affect a species in more subtle ways (e.g., behavioral changes versus immediate mortality) and often involve more indirect mechanisms. The response time is longer, and distinguishing stress-induced responses from normal environmental fluctuations and spatial heterogeneities is especially difficult. Temporal relationships, such as the one between biodegradation and accumulation rates, become important. Even more difficult to understand are synergisms, in which chronic pollutants have a greater impact in combination than separately. Each of these aspects contributes to the degree of uncertainty inherent in ecological evaluations. The significance of the uncertainty that remains after the ecosystems have been realistically characterized is a major issue in selecting among disposal options. This is particularly true since the uncertainty in predicting ecosystem-level effects of waste disposal includes not only a component related to the amount of information that has been collected about a system but also a component of intrinsic unpredictability. Further, the degree of uncertainty remaining even after a system is reasonably well studied varies from one ecosystem type to another.

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135 ~ 11 ~.. U]o us ~ in ~ ~ ~C .. en ~ ~o ~- o ~ ~- v ~3 US UP ~ o ~- - a, ~U] 0 ~a, o - ~ a) ~In ~ O o o - a ~ ~ ~ - ~1 ~ 11 ~a) ~ ~ Q C ~- O - ~ =,' i- , ~ - Up ~ Up ~ ~ ~- - Q Q ~ O - O ~ - 0) 0 - ~- Q Q ~ ~ ~O ~O O In us ~ ~ in O O ~-- V-- ~ ~ ~' - Q. ~Q In ~ ~ ~ ~C ~.- O O ~~-- vt ~- - ~ 3 G) ~ ~3 S ~- o i4 ,' U] ~ ~ . ~2 11 lU a) v' ~- ~- ~ ,.~: ~ ~ ~ =: a) ~ ~n me U~ ~ ~-1 Ln ~ ~ ~0 ~ 3 ~ - ' ~ ~Q e . -= ,' 5: ~ ~, ~ ~ ~.- e. U] O ~ - ~U] ~ ~- 11 ~- -- - ~ln ~ ~ ~ ., ~ ~ ~ 8 o ~ 3 ~ ~ U ,1 ~ ~ ~ ~ O O O C a. - ~ ~ GS r~ O ~1 - U] V] ~ ~- U, ~) ~ ~ O O I U] ~ ~ ~ ~ 1 ~ ~1 X O ~ I ~- ~ 1 - 11 4, ~ ~ 1 3 Q -v' _ u? ~n I c 'l ~ O ~ ~ 0 qJ ~ ~c o~ P ~ O ~ ~ - 11 ~ 11 ~ ~ ~ ~ 1 U] O ~ U] ~ ~ ~ ~ ~C ~ ~ ~ ~ C O ~ ~ ~ ~- - ~ - ~ . ~ . t, ~ 0 ~ ~ ~ ~ ~ ~ Q~ ~ ~- ~ . O ~ Q~ ~ ~ ~ ~ ~ O ~ ~ ~ ~ ~ O ~ ~ ~g ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ O O :~: % ~ ~- ~n v ~u~ ,1 .,, 0e .,.

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136 The effects of incomplete degradation or metabolic alteration of toxic chemicals on human health and ecosystems continue to be a subject of primary concern owing to the mutagenic, carcinogenic, and teratogenic potential of toxic parent compounds and their metabolites. REFERENCES Alderslade, R. 1981. The problems of assessing possible hazards to the public health associated with the disposal of sewage sludge to land. In Recent Experience in the United Kingdom--Characterization, Treatment and Use of Sewage Sludge, P. L. Hermite and H. Ott, eds. Reidel Publ. Co., Amsterdam, Holland, pp. 372-388. Allen, M. J. 1975. Regional variation in the structure of fish communities. In Annual Report, Southern California Coastal Water Research Project, E1 Segundo, Calif., pp. 99-102. Atwood, D., D. W. Brown, V. Cabelli, J. Farrington, C. Garside, G. Han, D. V. Hansen, G. Harvey, K. S. Kamlet, J. O'Connor, L. Swanson, D. Swift, J. Thomas, J. Walsh, and T. Whitledge. 1979. The New York Bight. In Assimilative Capacity of U.S. Coastal Waters for Pollutants, E. Goldberg, ed. U.S. Department of Commerce, NOAA Environmental Research Laboratories, Boulder, Colo., pp. 148-178. Bakelaar, R., and E. Odum. 1978. Community and population level responses to fertilization in an old-field ecosystem. Ecology 59:660-671. Baron, R. C., F. D. Murphy, H. B. Greenberg, C. E. Davis, D. J. Bregman, G. W. Gary, J. M. Hughes, and L. B. Schonberger. 1982. Norwalk gastroenteritis illness, Am. J. Epidemiol. 115:163-172. Bascom, W. 1982. The effects of waste disposal on the coastal waters of southern California. Environ. Sci. Technol. 16(4):226A-236A. Bastian, R. K. 1981. EPA's role and interest in using wetlands for wastewater treatment. Paper presented at the Midwest Conference on Wetland Values and Management, St. Paul, Minn., June 17-19, 1981. Bergh, A. K., and R. S. Peoples. 1977. Distribution of polychlorinated biphenyls in a municipal wastewater treatment plant and environs. Sci. Total Environ. 8:197-204.

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137 Boehm, P. D., S. Drew, T. Dorsey, J. Yarko, N. Moseman, A. Jeffries, D. Pilson, and D. Fiest. In press. Organic pollutants in New York Bight suspended particulates: waste deposits versus riverine/estuarine sources. In Wastes in the Ocean. Vol. 6. Near-shore .. . Waste Disposal. B. H. Ketchum, J. M. Capuzzo, I. W. Deudall, W. V. Burt, P. K. Park, and D. R. Kester, eds. Wiley-Interscience, New York. Boesch, D. F. 1982. Ecosystem consequences of alterations of benthic community structure and function in the New York Bight region. In Ecological Stress and the New York Bight: Science and Management, G. F. Mayer, ed. Estuarine Research Federation, Columbia, S. Carolina, pp. 543-568. Bopp, R. F., H. J. Simpson, C. R. Olsen, R. M. Trier, and N. Kostyk. 1982. Chlorinated hydrocarbons and radionuclide chronologies in sediments of the Hudson River and Estuary, New York. Environ. Sci. Technol. 16:666-675 Bouwer, H., J. C. Lance, and M. S. Riggs. 1974. High rate land treatment II. Water quality and economic aspects of the Flushing Meadows Project. J. Water Pollut. Control Fed. 46:844-859. Brooks, N. H., R. G. Arnold, R. C. Y. Koh, G. A. Jackson, and W. K. Faisst. 1982. Deep Ocean Disposal of Sewage Sludge off Orange County, California: A Research Plan. Environmental Quality Laboratory Rep. No. 21, Calif. Inst. Technol., Pasadena, Calif., 117 pp. Brown, D. A., R. W. Gossett, and K. D. Jenkins. 1982. Contaminants in white croakers Genyonemus lineatus (Ayres, 1855) from the southern California Bight: II. Chlorinated hydrocarbon detoxification/toxification. In Physiological Mechanisms of Marine Pollutant Toxicity. W. B. Vernberg, A. Calabrese, F. P. Thurberg, and F. J. Vernberg, eds. Academic, New York, pp. 197-213. Bryan, F. L. 1980. Epidemiology of foodborne diseases transmitted by fish, shellfish and marine crustaceans in the United States 1970-1978. J. Food Protection 43:859. Bryan, G. W. 1979. Bioaccumulation of marine pollutants. Phil. Trans. R. Soc. London B 286:483-505. Bryan, G. W. 1980. Recent trends in research on heavy metal contamination in the sea. Helogolander Meeresunters 33:6-25. Cabelli, V. J. 1978. Swimming-associated disease outbreaks. J. Water Pollut. Control Fed. 50:1374-1377.

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138 Cabelli, V. J. 1980. Health effects criteria for marine recreational waters. U.S. Environmental Protection Agency, Washington, D.C. Cabelli, V. J., A. P. Dufour, L. 3. McCabe, and M. A. Levin. 1982. Swimming associated gastroenteritis and water quality. Am. J. Epidemiol. 115:606. Califano, R. J. 1981. Accumulation and tissue distribution of polychlorinated biphenyls (PCBs) early life stages of the striped bass, Morone saxatilis. Ph.D. Dissertation. New York University, New York, 163 pp. In Capuzzo, J. M., and B. A Lancaster. In press. Responses of zooplankton populations to industrial wastes discharged at Deepwater Dumpsite 106 and their usefulness in estimating assimilative capacity. In Wastes in the Ocean, Volume 5 - DeePwater Disposal. W. , Burt, D. R. Kester, I. Duedall, P. K. Park, and J. M. Capuzzo, eds. Wiley Interscience, New York. Chaney, R. L. 1980. Health risks associated with toxic metals in municipal sludge. In Sludge--Health Risks of Land Application, G. Bitton et al., eds. Ann Arbor Science Publishers, Inc., Ann Arbor, Mich., pp. 59-83. Chaney, R. L. 1982. Foodchain pathways for toxic metals and toxic organics in wastes, Manuscript, 50 pp. Collett, J. N., and D. L. Harrison. 1968. Lindane residues on pasture and in the fat of sheep grazing pasture treated with lindane spills. N.Z. J. Agr. Res. 11:589-600. Colwell, R. R. 1980. Human pathogens in the environment. In Microbiolo~Y--1980. D. Schlessinger, ed. Am. Soc. for Microbiol., Wasb~ngton, D.~., pp. 337-379. Cook, C. W. 1976. Surface-mine rehabilitation in the American west. Environ. Conserv. 3:179-183. Council for Agricultural Science and Technology. 1976. Application of sewage sludge to cropland. U.S. EPA, Office of Water Programs, Washington, D.C. EPA-430/9-76-013. Duedall, I. W., H. B. O' Connors, S. A. Oakley, and H. M. Stanford. 1977. Short-term water column perturbations due to sewage sludge dumping in the New York Bight Apex. J. Water Pollut. Control Fed. 49:2074-2080. Ewel, K. C., M. A. Harwell, J. R. Kelly, H. D. Grover, and B. L. Bedford. 1982. Evaluation of the use of natural ecosystems for wastewater treatment. Ecosystem Research Center Rep. No. 1S, Cornell University, Ithaca, N.Y., 55 pp.

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141 Loehr, R. C., W. J. Jewell, J. D. Novak, W. W. Clarkson, and G. S. Friedman. 1979. Land Application of Wastes. Van Nostrand-Reinhold Co., New York, 431 pp. Malins, D. C., and T. K. Collier. 1981. Xenobiotic interactions in aquatic organisms: effects on biological systems. Aquat. Toxicol. 1:257-268. Malins, D. C., and B. B. McCain, D. W. Brown, A. K. Sparks, and H. O. Hodgins. 1980. Chemical contaminants and biological abnormalities in central and southern Puget Sound. NOAA Tech. Memo. OMPA 2-, 295 pp. McCain, B. B., H. O. Hodgins, W. D. Gronlund, J. W. Hawkes, D. W. Brown, M. S. Myers, and J. H. Vandermuelen. 1978. Bioavailability of crude oil from experimentally oiled sediments to English sole (Parophrys vetulus), and pathological consequences. J. Fish. Res. Bd. Can. 35:657. McIntosh, R. P. 1980. The relationship between succession and the recovery process in ecosystems. In The Recovery Process in Damaged Ecosystems, J. Cairns, ed. Ann Arbor Science, Ann Arbor, Mich., pp. 11-62. Mearns, A. J. 1981. Ecological effects of ocean sewage outfalls: observations and lessons. Oceanus 24(1):45-59. Mearns, A. J., and J. Q. Word. 1982. Forecasting effects of sewage solids on marine benthic communities, pp. 495-512 in Ecological Stress and the New York Bight: Science and Management. G. F. Mayer, ed. Estuarine Research Federation, Columbia, S. Carolina, pp. 495-512. Mearns, A. J., E. Haines, G. S.Kleppel, R. A. McGrath, J. J. A. McLaughlin, D. A. Segar, J. H. Sharp, J. J. Walsh, J. Q. Word, D. K. Young, and M. W. Young. 1982 Effects of nutrients and carbon loadings on communities and ecosystems. In Ecological Stress and the New York Bight: Science and Management, G. F. Mayer. ed. Estuarine Research Federation, Columbia, S. Carolina, pp. 53-65. Metcalf, T., and W. Stiles. 1968. Viral pollution of shellfish in estuary waters. J. Sanitation Eng. Div. ASCE, 94:595. Milton, W. 1947. The yield, botanical and chemical composition of natural hill herbage under manuring, controlled grazing and hay conditions. I. Yield and botanical. J. Ecol. 35:65-89. Mitchell, R., and C. Chamberlain. 1979. Survival of indicator organisms. In Indicators of Viruses in Water and Food, G. Berg, ed. Ann Arbor Science Publishers Inc., Ann Arbor, Mich.

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