Methods
Stress Index = 1.0-(N + I),
where N +1 varies from 1 (total destruction) to 0 (zero mortality).
Although there are problems in applying this index in the field, development of a
quantitative index of stress would assist in the assessment of natural versus
anthropogenic contributions in hard-bottom benthic communities (Brown, 1988).
Indicator Species
Examination of abundances of individual indicator species are generally informative and
may reduce the cost of the analysis. The absence of pollution-sensitive species and the
enhancement of opportunistic and pollution-tolerant species may assist in defining the
spatial and temporal extent and magnitude of impacts. However, indicator variables
must possess the following characteristics (Green, 1979):
• Must provide a sufficiently precise and accurate appraisal of:
- species of concern,
- anthropogenic disturbances to benthic communities, and
- presence/absence or magnitude of anthropogenic perturbation to the
ecosystem.
• Must provide a cost-effective and statistically reliable alternative to monitoring
all critical benthic community measures of habitat perturbation.
• Must be appropriate for the spatial and temporal scale demanded by the
study objectives.
Pearson and Rosenberg (1978) demonstrated that along a gradient of organic
enrichment, a predictable community of benthic infauna could be observed (Figure 4-3).
Communities consist of opportunistic, tolerant species in areas of severe pollution,
giving way to less tolerant and more competitively dominant species further from
severely polluted areas (Figure 4-3). This model of indicator species composition and
distribution has been found to be useful in assessing both natural and anthropogenic
disturbances.
Several indices, derived from the distributions of pollution-sensitive species and
opportunistic disturbance-tolerant species, have been developed to evaluate impacts to
infaunal benthic community structures: copepod/nematode ratio (Raffaelli and Mason,
1981), infaunal trophic index (Word, 1978), and organism-sediment index (Rhoads and
Germane, 1986). The infaunal index, which was initially developed for the Southern
California Bight, uses the abundances of four functional groupings to describe
community structure. Low values of the infaunal index indicate communities dominated
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Figure 4-3. Diagram of Changes in Fauna and Sediment Structure Along a Gradient of
Organic Enrichment (from Pearson and Rosenberg, 1978)
by deposit-feeding organisms that are considered to be more impacted than
communities dominated by suspension feeders (high infaunal index values)(Word,
1978).
Polluted or disturbed conditions may be indicated by:
• High rates of the nematode/copepod ratio;
• Low values (A) of the infaunal index; and
• Low values (7) of the organism-sediment index.
However, further studies of the response patterns of infaunal species subjected to
anthropogenic perturbations are required in order to select appropriate indicators of
benthic community impact. Again for hard-bottom communities, the use of epifaunal
indicator species and other indices requires further research (Brown, 1988).
It is recommended that multiple pollution-tolerant and pollution-sensitive indicator
species be selected by clearly defined criteria. Multiple indicator species often provide a
more complete representation of environmental conditions.
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Multivariate Analyses
Numerical classification encompasses a wide variety of techniques that have been used in
the analysis of benthic data to distinguish groups of entities (e.g., sample locations)
according to similarity of attributes (e.g., species). These techniques differ from most
multivariate methods in that no assumptions are made concerning the underlying
distributions of the variables. Detailed descriptions of numerical classification analysis can
be found in Pielou (1984), Romesburg (1984), Clifford and Stephenson (1975), Boesch
(1977), Sneath and Sokal (1973), and Anderberg (1973). Boesch (1977) is particularly
valuable as an introduction and guide to the use of numerical classification analysis in
marine environmental studies. Guidance on the interpretation of classification results is
provided in an EPA Technical Support Document (USEPA, 1988d).
Ordination analyses have also been used to reduce the dimensionality of the data set while
maintaining the relationship among similar and dissimilar entities. At present, no single
ordination technique has been shown to be clearly superior for the analysis of biological data
(USEPA, 1985d).
Multivariate analyses are effective heuristic tools. They generate visual representations
that often indicate where further analyses ought to be conducted.
Analytical Approach Recommendations
Some of the most informative measures of community structure are the simplest (Table
4-20):
• Number of individuals,
• Number of species,
• Dominance,
• Infaunal index,
• Abundances of pollution-sensitive species, and
• Abundance of opportunistic and pollution-tolerant species.
These indices have proved to be useful over various habitats and regions in assessing
changes to benthic community structures (USEPA, 1985d). Values of these indices may
be determined from the list of species abundances generated during the taxonomic
identifications of collected specimens. Furthermore, the values of these six variables may
be easily tested statistically using parametric or nonparametric techniques. It is
recommended that no single index or analytical method be used to assess impacts;
instead, the assessment of impacts should incorporate information that each variable
and method contributes concerning benthic community structure.
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Selection of reference sites is key to the evaluation of environmental impact
assessment. Results of analyses using reference measures provide the means of
comparison by which anthropogenic impacts are detected. It is essential that selected
reference sites exhibit at least similar:
• Sediment characteristics (i.e., grain size or substratum type),
• Water depths,
• Flow characteristics,
• Salinity,
• Dissolved oxygen, and
• Temperature
compared to monitoring program sampling sites. Several reference sites may be
required to meet these criteria.
4.5.4 QA/QC Considerations
Sample Collection
Surveying and sampling equipment should be inspected for wear and tear to avoid loss
of data or sample leakage and loss upon ascent. It is recommended that backup survey
and sampling equipment be available on board the vessel in case the primary
equipment breaks down or is lost during the cruise.
The following infauna sample acceptability criteria should be satisfied (USEPA,
1986-1991):
• Sediment is not extruded from the upper face of the sampler such that
organisms may have been lost.
• Overlying water is present, indicating minimal leakage.
• The sediment surface is relatively flat, indicating minimal disturbance or
winnowing (Figure 4-4).
• The entire surface of the sample is included in the sampler.
• The desired penetration depth is achieved.
If the sample does not meet all the criteria, it should be rejected.
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Acceptable if Minimum
Penetration Requirement Met
and Overlying Water is Present
Unacceptable-
(Canted with Partial Sample)
Unacceptable
(Washed, Rock Caught in Jaws)
Unacceptable
(Washed)
Figure 4-4. Examples of Acceptable and Unacceptable Samples (USEPA, 1987c)
Appropriate QA/QC protocols for assessing the validity of hard-bottom surveys (e.g.,
multiple transects, resurveying one or more transects or quadrats by different divers)
should also be followed.
Taxonomic Identification
A key QA/QC issue is taxonomic standardization. Consistent taxonomic identifications
are achieved through interaction among taxonomists working on each major group.
Participation of the laboratory staff in regional taxonomic standardization programs is
recommended to ensure regional consistency and accuracy of identifications.
Five percent of all samples identified by one taxonomist should be reidentified by
another taxonomist who is also qualified to identify organisms in that major group.
It is advisable that at least three individuals of each taxon should be sent for verification
to recognized experts. These verified specimens should then be placed in a permanent
reference collection. All specimens in the reference collection should be stored in
labeled vials that are segregated by species and sample. Reference specimens should
be archived alphabetically within major taxonomic groups.
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It is also recommended that at least 20 percent of each sample be re-sorted for QA/QC
purposes. Re-sorting is the examination of a sample or subsample that has been sorted
once and is considered free of organisms. Re-sorting should be done by someone other
than the one who sorted the original sample. If a sample is found that does not meet the
recommended 95 percent removal criterion, the entire sample should be re-sorted. For
epifaunal surveys, a portion of the photographed quadrats should be reanalyzed for
comparisons.
4.5.5 Statistical Design Considerations
Consideration of statistical strategies will mitigate the high costs of collecting and
processing samples.
Temporal Stratification of the Data
The time of the year should be controlled or stratified in the design; the use of annual
averages is seldom a good practice. Temporal stratification of the data should not be
attempted until sufficient knowledge of long-term natural cycles is attained. Initially,
simple regression analyses may be conducted on seasonally stratified data to identify
monotonic temporal trends. Further examinations of whether conditions are improving
or degrading over time may be performed using various statistical time series analyses.
Statistical Power
Selection of the number of replicates is an important component of program design.
The inherent patchiness of benthic communities requires collection of sufficient replicate
samples to ensure an accurate description of the benthos. However, increases in
replication increase sample processing costs. Power analyses assist in the allocation of
sampling resources (stations, replication, and frequency) with regard to program
finances and design.
Power analyses may be applied to determine the appropriate number of sample
replicates required to detect a specified difference (USEPA, 1987d). The number of
replications required to detect a specified minimum difference is a function of the
statistical power and the variance in the data. Power analyses require a priori
knowledge of the variability in the data. A best guess or, preferably, variation observed
in historical data is often used initially in the design of the monitoring program.
To improve the power of a statistical test, while keeping the significance level constant,
the sample size (area sampled, number of replicates of grabs, quadrats, or transects)
should be increased. Because of constraints in cost and time, however, this option may
not be available. Power analysis has shown that for a fixed level of sampling effort, a
monitoring program's power is generally increased by collecting more replicates at fewer
locations. Sampling should be conducted in a radiating pattern from the zone of initial
discharge out to the distance where there will be no effects.
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4.5.6
Use of Data
Monitoring of benthic community structure provides in situ measures of the benthic
habitat and is a powerful tool in the evaluation of spatial and temporal effects of
anthropogenic and natural disturbance. The presence or absence of certain organisms
is useful in indicating the previous condition of the environment (Bilyard, 1987; Tomascik
and Saunder, 1987). Monitoring of benthic infaunal communities also provides data
required in the design and validation of benthic community dynamics models (Pearson
and Rosenberg, 1978; Brown, 1988) and the selection of biological indicators (Word,
1978).
In addition, monitoring of benthic communities directly provides accurate information
essential in assessing the effectiveness of discharge reductions (Bilyard, 1987). For
example, benthic infauna monitoring provided information used to assess the
effectiveness of pollution abatement plans in the recovery of Southern California waters
(Reish, 1986; SCCWRP, 1988). These studies indicate that analyses of benthic
communities may be effectively used to monitor the long-term health of the receiving
environment (Reish, 1986).
i
Currently, benthic communities have not proved useful for identifying specific chemicals
or classes of chemicals present in the environment. Further information concerning
specific responses to specific contaminants is required before infaunal community
structure becomes useful in identifying specific contaminants (USEPA, 1989e). In
addition, caution is recommended in the use of benthic community structure to predict
specific effects on potential predators. Information on trophic relationships, competition,
and predation is often not available. The capability to predict the effects of altered prey
communities on predators may improve with research on these topics. Factors such as
prey quality, distribution of prey, and interactions among species will be important
components of this research.
However, benthic invertebrates do serve as effective indicators of environmental
condition, delineating the magnitude, spatial extent, and temporal trends of
anthropogenic and natural perturbations to the ecosystem (Reish, 1986; Bilyard, 1987).
Monitoring of benthic infauna and epifauna will provide relevant accurate data
fundamental to achieving the objectives of most monitoring programs.
4.5.7 Summary and Recommendations
Rationale
The objective is to detect and describe spatial and temporal changes in the
structure and function of benthic communities.
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• Benthic monitoring provides in situ measures of habitat quality and is a
powerful tool in assessing environmental impact.
Monitoring Design Considerations
• It is recommended that consistent types of surveying and sampling gear, data
collection and sample sorting protocols, level of taxonomy, and location and
timing of sample collection be implemented to allow for comparisons between
studies.
• To reduce the variation due to seasonal differences, sampling should be
conducted during the same season—preferably the same month—each year.
• Voucher specimens should be collected for in situ epifaunal surveys.
• For epifaunal surveys, data are collected on species and organism coverage
of the substratum by diver, ROV, or submarine from specified transects or
quadrats.
• For infaunal sampling, collection of undisturbed sediment requires that the
sampler:
- create a minimal bow wake when descending;
- form a leakproof seal when the sediment sample is taken;
- prevent winnowing and excessive sample disturbance when ascending; and
- allow easy access to the sample surface in order that undisturbed
subsamples may be taken.
• Penetration well below the desired sampling depth is preferred to prevent
soft-bottom disturbance as the device closes.
• Grab samplers and box corers are recognized as the tools of choice for
maximum accuracy and precision when sampling soft-bottom habitats.
• Sorting through a standard sieve mesh size (i.e., 0.5 mm) is recommended.
Further sorting through other mesh sizes may be conducted in addition to
sorting through this standard mesh size.
• Relaxants facilitate identification and morphometric measurements; however,
standard procedures must be implemented to ensure valid comparisons
among studies.
• Vital stains may facilitate sorting; however, a proper QA program should
ensure that sorting efficiency is maintained.
• Identifications to higher taxonomic levels may be sufficient to meet program
objectives; however, it is recommended that all samples be archived if
comparisons to lower taxonomic levels will be required at a later date.
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Analytical Methods Considerations
• It is recommended that simple measures of community structure be used to
assess the condition of the benthos: number of individuals, number of
species, dominance, infaunal index, abundance of pollution-sensitive
species, abundance of pollution-tolerant species, and percent cover for
hard-bottom indicator groups.
• Selected biological indices should retain biological meaning, be robust
indicators of ecosystem "health," and incorporate species form and function.
• Indicator species should possess the following characteristics:
- sensitive to benthic perturbations of concern;
- cost-effective and statistically reliable alternative to measuring all species
in a monitoring program;
- statistically reliable indicative measures of habitat perturbations; and
- appropriate for the spatial and temporal scale demanded by the study
objectives.
• Selection of reference sites is key to the evaluation of environmental impact
due to anthropogenic impacts; several reference sites may be required.
QA/QC Considerations
• Taxonomic standardization is essential to the analysis of community
structure. Recommended protocols include consistent interactions among
taxonomists, reidentification of selected samples, use of a reference
collection, and re-sorting and analysis of selected samples or subsamples.
Statistical Design Considerations
• Power analyses may be applied to determine the appropriate number of
sample replicates required to detect a specified difference, thereby optimizing
the high costs of collecting and processing samples.
Use of Data
• Data provide essential information in order to assess impacts due to
anthropogenic perturbation, monitor recovery of the receiving environment,
and validate community and population models.
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4.6
FISH AND SHELLFISH PATHOBIOLOGY
Pathobiological methods provide information concerning damage to organ systems of
fish and shellfish through an evaluation of their structure, activity, and function.
Anatomic pathology methods can give an indication of the nature of an altered state, for
example, by identifying the specific type of tumor present in an animal. Reproductive
developmental studies examine the reproductive capacity of animals and can provide
information to aid in estimating and predicting population abundance and recruitment.
Biochemical/enzymological studies seek to detect differences in enzymatic activity as a
measure of biological condition. Immunological methods can demonstrate altered
immune response, an indicator of changes in bodily defense mechanisms and increased
susceptibility to disease.
Pathobiological methods should be used in concert to investigate cause-and-effect
relationships as a result of contaminant exposure. Anatomic pathology can serve as a
vital link between observed effects on populations and communities in an estuary and
the changes in activity and function observed by other methods.
4.6.1
Rationale
Pathobiological methods can be used to examine adverse effects of pollutants on fish
and shellfish. The presence of toxics in water and sediments may not immediately
result in visible changes in these organisms. Biomarkers offer a more sensitive and
reliable assessment of exposure risks than ambient water or sediment quality
monitoring. Monitoring of pathobiological effects provides information necessary to
make determinations of the existence of adverse effects in animals (e.g., tumors),
population productivity and stability (affected by reproduction and disease states), and
the loss of organisms deemed valuable for ecological, aesthetic, recreational, scientific,
or economic reasons. Table 4-21 outlines some of the terms used to describe
pathobiological methods.
Although the value of these methods for establishing cause-and-effect links has been
established during laboratory toxicity studies, some questions remain regarding their ability
to establish such links for field-collected organisms that are exposed to a variety of natural
environmental stresses and combinations of contaminants (Hinton and Couch, 1984; Couch
and Harshbarger, 1985; Mix, 1986; Sindermann, 1990). However, properly conducted
multidisciplinary monitoring studies using these methods can provide regulatory agencies
with evidence of impaired health status in animals exposed to contaminants in estuarine
ecosystems. This information can then be used to direct laboratory confirmation of the
cause, if necessary (see, for example, Buckley etal., 1985; Gardner et al., 1991). Continued
monitoring with these methods can be used to detect changes in a population's health
during and following environmental intervention. Because changes at the organismal
level precede changes in population and community characteristics, pathobiological
studies can provide an early indication of the effectiveness of management actions.
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biomarker
cytochrome P450
cytogenetics
genotoxic agents
hepatic
histopathology
hfstochemistry
immunoassay
immunology
inclusion bodies
macrophage
pathobiology
smooth endoplasmic
reticulum (SER)
Table 4-21. List of Pathobiological Terms
any biological method used to detect the exposure of organisms to hazardous
chemicals in the environment by measuring the response of the organisms to the
contaminant through comparative molecular, biochemical, physiological, or
anatomical observations of cellular dysfunction
a protein in the microsomes of liver cells that is important in catalyzing the
metabolism of steroid hormones and fatty acids and in the detoxification of a variety
of chemical substances
the study of cytology in relation to genetics, especially the study of chromosomal
behavior in mitosis and meiosis. Modern cytogenetics has led to the identification of
chromosomes as bearers of the genes and deoxyribonucleic acid (DMA) as the key
molecule of the gene
chemical and physical agents that can produce genetic alterations at subtoxic
concentrations and can result in altered heredity characteristics. Genotoxic agents
generally possess specified chemical or physical properties that facilitate their
interaction with nucleic acids
of, or relating to, the liver
pathologic histology; the science or study dealing with the cytologic and histologic
(microscopic) structure of abnormal or diseased cells, tissues, and organs in relation
to their function
the study of the chemistry of cells and tissues using light and electron microscopy,
special chemical tests, and special stains to determine the location of certain enzyme
systems or reaction products in the cell
measuring the protein and protein-bound molecules that are concerned with the
reaction of an antigen with its specific antibody, as in the detection of hormones or
other substances
the study of being protected from a disease; the study of the response of the body
and its tissues to a variety of antigens, including red cells, pollens, transplanted
tissues, and even the individual's own cells
bodies present in the nucleus or cytoplasm of certain cells in cases of infection by
filtrable viruses or as the result of degenerative diseases or exposure to chemicals
cells of the reticuloendothelial system having the ability to phagocytose particulate
substances and to store vital dyes and other colloidal substances
pathology, the study of disease, with emphasis more on the biological than the
medical aspects of the essential nature, causes, and development of abnormal
conditions, as well as the structural and functional changes that result from the
disease process
a connecting network of tubules that course through the cytoplasm of the cell and can
be viewed using the electron microscope; SER is essential to metabolic functions of
the cells
SOURCES: Stedman's Medical Dictionary, 1982; Taber"s Cyclopedic Medical Dictionary, 1985.
EL'
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mSA
•:ir
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Section 403 Procedural and Monitoring Guidance
4.6.2 Monitoring Design Considerations
A field survey to collect target organisms and tissue samples may be required for
pathobiological monitoring (Hargis et al., 1984; U.S. EPA, 1987b). In certain instances,
a large sample size may be needed to establish statistical significance because of
normal variation from animal to animal, species and generic differences, and migratory
habits of fish. For these reasons, and the labor-intensive nature of pathobiological
methods, the availability and allocation of funds, time, equipment, and trained personnel
must be considered when planning to include pathobiological methods in monitoring
programs.
Selection of Target (Indicator) Species
A key component of any pathobiological monitoring program is the selection of target
species. The fundamental criterion is the ability to use the selected species to make
comparisons between sampling locations and sampling periods. It is recommended that
the target species possess the following characteristics:
• Abundant enough, temporally and spatially, to allow for adequate sampling;
• Large enough to provide adequate amounts of tissue for analysis;
• Sedentary (nonmigratory) in nature to ensure that pathobiological
abnormalities are representative of the study area; and
• Easily collected.
It is also recommended that an existing database documenting exposures and
sensitivities of the target organism specific to contaminants of concern is available.
Fish provide sufficient tissue for analyses and may indicate potential threats to human
populations. However, fish are motile and pathobiological abnormalities detected may
not be representative of the study area (see Vogelbein et al., 1990).
Bivalve molluscs, either attached to the substratum or burrowing in sediments, have
been sampled extensively to examine the condition of these commercially important
estuarine species (e.g., Mussel Watch, National Status and Trends program). The most
common target species have been oysters (Crassostrea virginica) and mussels (Mytilus
spp.), although pathobiological and bioassay studies have also been performed on other
species, such as crabs and penaeid shrimp.
Secondary considerations, based on economic importance and status as a bioassay
organism, may be applied to further winnow the list of candidate target species to a
practical number of species to be analyzed.
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Sampling Location
Appropriate locations of sampling stations depend on the objectives of the study. For
example, to evaluate whether there is a statistically significant increase in lesions,
stations should be located to collect specimens from contaminated and uncontaminated
(background or control) areas for statistical comparison. It is also important to
demonstrate a dose-response relationship between the pollutant concentration and
incidence of tumors or other lesions by locating sample stations along a contamination
gradient (i.e., from highly contaminated to moderately contaminated to uncontaminated).
It is recommended that stations be located in areas where the geographic area of
contamination is large enough that sampled fish could reasonably be expected to have
spent a considerable amount of time within the influence of the pollutant (U.S. EPA,
1987b). Nielsen and Johnson (1984) and Cailliet et al. (1986) may be consulted for
further fish sampling methodologies. Information on sampling bivalves and other
invertebrates is contained in Couch (1978), Yevich and Barszcz (1983), Turgeon et al.
(1991), and other sources.
4.6.3 Analytical Methods Considerations
Anatomic Pathology Methods
Anatomic pathology methods examine tissues with the naked eye (gross anatomic
pathology), the aid of a light microscope (LM methods), or the electron microscope (EM
methods). Gross anatomic methods are concerned with obvious adverse changes in
tissue that can be observed in the field (Hunn, 1988; Hargis et al., 1984). The
advantage of gross methods is that large numbers of specimens can be examined
rapidly. However, the methods are generally nonspecific (i.e., it is not possible to
determine that a specific pollutant led to a specific disease). An exception to this is
cataracts in fish, which have been linked with polynuclear aromatic hydrocarbon
pollution in the field (Hargis and Zwerner, 1988).
A disadvantage of anatomic pathology methods is that they require specialized
personnel and laboratories. The methods are, however, generally standardized, routine,
and operational in existing Federal, State, university, veterinary, and private diagnostic
laboratories that specialize in aquatic animal pathology. Current activities are aimed at
developing a field-to-laboratory response-diagnostic scenario in which adverse effects
are found in the field, diagnosis is made in the laboratory, and experimental studies are
initiated to verify results. Refer to Yevich and Barszcz (1980), Howard and Smith
(1983), Johnson and Bergman (1984), Klontz (1985), Meyers and Hendricks (1985), and
U.S. EPA (1987b) for more information on these methods and techniques.
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Light Microscopy Methods
LM (histologic) methods use the light microscope to view cells and tissues. These
methods can detect changes such as inflammatory responses, alterations in the
appearance of cells, cancerous/precancerous lesions, and damage due to parasites
(USEPA, 1986c). The advantages of LM methods are that they are organ-specific (i.e.,
the lesion can be localized to a specific organ) and can detect microscopic and subtle
cellular alterations that are not evident on visual inspection. However, LM methods are
more expensive (approximately $30/sample) and slower than gross methods.
Electron Microscopy Methods
EM methods use the electron microscope to detect changes in tissue at cellular and
subcellular levels, such as the identification of the nature of inclusion bodies or changes
in the amount of smooth endoplasmic reticulum. An advantage of EM methods is that
they can be highly specific because the pollutant can be localized within certain parts of
the cell. These inclusions may contain the causative chemical agent; lead, gold, iron,
bismuth, uranium, beryllium, mercury, copper, and arsenic are a few of the metals that
can be deposited intracellularly (Sorenson and Smith, 1981). Additionally, EM methods
can be used to investigate subcellular mechanisms of pollutant action. The major
disadvantage of EM methods is that they are very expensive (minimum $400/fish) and
very slow, requiring highly skilled technical expertise.
Histochemical Methods
Histochemical methods use the microscope in conjunction with special chemical tests
and stains to localize specific enzyme systems or reaction products in the cell. For
instance, histochemical assays have been used on liver tissue during and after tumor
formation to yield important information on the biochemistry of specific lesions (Hinton et
al., 1988; Prophet et al., 1992; Sumner, 1988). Histochemical methods are reliable and
can be highly specific for certain classes of organic compounds and metals. However,
highly skilled technical expertise is required to carry out the methods in the laboratory.
In Vitro Tests
In vitro tests are generally more sensitive than whole animal systems, less expensive to
carry out, and of shorter duration. However, in in vitro tests, defense mechanisms found
in the intact animal are missing. In vitro systems offer the greatest flexibility for the
testing and study of environmental contaminants. They can be designed to be relevant
to the species of interest in a given area, and multiple types of measurements can be
taken from a single test system (e.g., metabolic products, mitotic activity, cytotoxicity
and genetic damage). However, comparative in vitro and in vivo studies are needed to
correlate and relate changes that occur in each system (Landolt and Kocan, 1983).
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Reproductive/Developmental Methods
Reproduction studies are designed to examine a number of parameters in the
reproductive process: gross examination of the egg and numbers of eggs, embryo
viability, the proportion of fertilized eggs (i.e., fertilization success), and larval
development and viability (USEPA, 1986c). Studies that examine the egg itself
incorporate microscopic methods that are time-consuming and expensive (West, 1990),
but serve as a direct measure of reproductive success. Some methods deal with the
egg at the molecular level and analyze the action of chemical and physical agents
whose toxicity is directed toward genetic (DNA) components of the egg.
Different cytogenotoxic tests can be used to measure a diverse array of effects including
gene mutation, chromosome damage (sister chromatid exchange), primary DNA
damage, or oncogenesis (i.e., tumor formation and development) (Brusick, 1980;
Landolt and Kocan, 1983; Klingerman, 1982; Shugart, 1990). Another area that is
currently being investigated as a suitable, sublethal assay for toxicity of certain wastes in
the marine environment is the fin regeneration test (Weis et al., 1990). The information
collected using these methods can be used to assess the effects of pollutants on the
reproductive capacity of animals. An understanding of these effects can be useful in
evaluating observed population and community-level changes relative to the occurrence
of specific pollutants.
Gonadotropic and steroidogenic hormones regulate the reproductive capacity of an
organism. The level of these hormones has been used to assess how pollutants affect
the reproductive capacity of fish (V. Varanasi, 1990, National Marine Fisheries Service,
NOAA, personal communication). The analysis is a sensitive indicator of exposure
affecting major biological processes that impact the whole population. However,
relatively detailed information about the normal reproductive cycle of the animals is
necessary to apply these methods in the field (V. Varanasi, 1990, National Marine
Fisheries Service, NOAA, personal communication).
Biochemical Methods
Biochemical methods have been used in field studies to measure various indicators of
environmental contamination. These methods are inherently sensitive and may provide
basic information about early changes in response to environmental contamination at
the cellular level. The development of a suite of indicators having both specific and
nonspecific responses can provide information on the type of stressors, mechanisms of
action, extent of physiological dysfunction, and potential long-term population
consequences (Thomas, 1990).
Fish can respond to generalized stress, contaminants being one type of stress, through
induction (increased synthesis) of stress proteins (Pickering, 1981; Sanders, 1990).
Stress proteins are currently being investigated for use as generalized biochemical
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Section 403 Procedural and Monitoring Guidance
indicators of stress in fish, chemical-class pollutant indicators, and mode of action
indicators. The methods for detecting stress proteins involve radioisotopic and
immunologic methods that measure the amount of stress protein present after a stress
(i.e., exposure to a pollutant) occurs. At present, cDNA probes are being used
experimentally to measure the correlation between stress and induction of stress
proteins. These methods potentially afford a high degree of sensitivity.
It has been suggested that induction of the fish hepatic microsomal mono-oxygenase
(MO) enzyme could serve as a sensitive biological indicator for certain classes of
chemicals in water (Payne et al., 1987; Kleinow et al., 1987; Lech et al., 1982; Jimenez
et al., 1990; Haux and Forlin, 1988). Metallothioneins (MT) have been under
consideration for use as a monitoring tool for trace environmental metal pollution due to
their induction as a result of exposure to certain metals (Engel and Roesijadi, 1987;
Garvey, 1990; Haux and Forlin, 1988). However, additional scientific research is
required to understand the basic biology of fish before the exact significance of field
studies using these techniques can be ascertained.
A concern when measuring biochemical variables in fish to detect environmental
pollutants is that their exact biological significance is rarely understood. In addition, for
most of the biochemical variables studied, the normal range for a particular fish
population and the factors influencing these variables are often unknown (Neff, 1985).
Even with these limitations, biochemical methods hold considerable promise as
sensitive early indices of exposure to environmental stressors (Thomas, 1990).
However, additional research is needed so that simplified, more cost-effective field
methodologies can be developed.
Immunoloaical Methods
Immunological biomarkers are simple, sensitive, reproducible, and workable in the field
(Weeks et al., 1990; D. Anderson, 1990; R.S. Anderson, 1990). These indicators
provide supportive evidence for linkage between a stressor (toxicant, etc.) and disease
outbreaks in fish and shellfish. The immune response can be used to monitor a specific
antigen or microorganism responsible for pathological conditions in fish. Biologists can
perform quick and sensitive assays in the field or in their own diagnostic laboratories
because many immune assays are becoming available in kits (Rowley, 1990; Matthews
et al., 1990). Many immunological assays do not require sacrifice of the animal. Blood
samples can be taken periodically to follow the kinetics of the effects of stress in a single
animal; however, the effects of handling stress on aquatic species must also be
considered in this case. The immune response is physiologically similar among most
vertebrates and similar equipment and materials can be used to test all species of fish
as well as shellfish (see Anderson, 1987). There is a rapidly growing body of literature
on immunotoxicology from veterinary and aquatic animal sciences.
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The selection of immune system indicators for the study of stress effects depends on
many factors, including specific study objectives, available equipment, training,
personnel, and length and number of assays. The most sophisticated and sensitive
assays are costly and require highly trained personnel, whereas simple assays can be
performed by field biologists with only basic laboratory supplies.
A major limitation of immune indicators is that the response is sometimes too broad to
provide conclusive evidence that the observed reaction is actually due to a specific
complex to be considered. Cross-reactions and heightened responses to nonspecific
factors may prevent the interpretation of assays with absolute certainty. The immune
response in fish or shellfish will be distinctive for a specific antigen or disease-causing
agent. This will frequently make it difficult to know which immune indicator is most
affected and which immunological assay to apply. Expensive materials and laboratory
equipment are needed for sophisticated immunological assays. Confirmatory assays
are advisable to be sure that a particular stressor is the only cause of a particular effect.
Many animals should be sampled because of natural variability among individuals.
Physiological Methods
Hematological methods have been used by biologists for many years to assess the
general health of fish in hatcheries and research laboratories. The procedures are
well-standardized (Blaxhall and Daisley, 1973; Wedemeyer and Yasutake, 1977) and
are easy to carry out, even in the field. Hematological measures may be affected by the
stress of capture, but are far less influenced than some other measurements, such as
blood glucose (Larson et al., 1985). Hematologic methods have been successfully
implemented in the field and have been rated as the best physiologic method for
evaluation of pollutants (U.S. EPA, 1986c). For the measurement of hematological
factors in fish, techniques similar to those used in human and veterinary clinical
laboratories are generally used with minor modifications (Heath, 1987; Bouck 1984).
Blaxhall and Daisley (1973), Wedemeyer and Yasutake (1977), and Ellis (1977) provide
practical guides to the adaptation of these methods for use on fish blood.
The hematocrii/erythrocyte determination may not be as sensitive to pollution as is the
leucocyte count, at least as far as its response to metals is concerned (Larson et al.,
1985). It remains to be determined how sensitive the procedure is to subtle
environmental changes. In general, chemical and physical stressors cause a decrease
in the leucocrit, whereas infections produce the opposite response. However, it is also
possible to obtain elevations in granulocytes concomitant with a decrease in
lymphocytes, thereby yielding an unchanged leucocrit (Peters et al., 1980). Thus, the
method has limitations for the detection of chronic stress (Wedemeyer et al., 1983).
t
Several obstacles, such as capture stress, limit the potential usefulness of physiological
tests in field work. Capture stress precludes the use of sensitive, early indicators of
environmental stress. The use of physiological responses of wild fish to assess
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environmental quality is difficult because responses due to toxicants often cannot be
distinguished from those induced during handling of the wild fish (Bouck, 1984; Folmar,
1993).
4.6.4 QA/QC Considerations
Pathobiological methods have a wide range of sensitivities and response times (i.e.,
when a response can be detected). For instance, certain immunologic, electron
microscopic, and biochemical methods can detect early changes in cells and are very
sensitive. Gross anatomic pathology methods, on the other hand, can detect cellular
changes only after lesions can be seen and, therefore, are less sensitive.
General considerations for QA/QC procedures have been covered earlier in this
document. With regard to pathobiological methods, it should be noted that careful and
consistent handling of aquatic specimens is required to minimize trauma and
confounding effects, such as exposure to air. Organisms should be held in the
laboratory under conditions as near to those found at the site of collection as possible
and for as short a time as practicable before performing assays. Fish and shellfish
collected for histopathological examination must be properly fixed (e.g., immersed in a
formaldehyde or glutaraldehyde solution) to stop metabolic activity. This may require
opening or sectioning the organisms to allow the fixative to rapidly penetrate all tissues
and preserve the cellular structure in its existing condition. The organisms must still be
active or moribund, but not dead, before being fixed. Failure to follow proper fixation
procedures will interfere with the interpretation of lesions in anatomic pathology studies.
As with other monitoring methods, samples must be accurately labeled at the time of
collection and routine QA/QC procedures should be instituted, including tracking
samples, carefully recording methods used, using fresh solutions, and treating both
control and exposed samples equally. Whenever possible, organisms for each group to
be tested should be of the same species, age, and sex. Sections of tissues and organs,
or for small organisms, the whole animal, should also be prepared as uniformly as
possible with respect to homogeneity and orientation so that microscopic observations
can be made on the same organs and areas for each specimen. Subsamples of
sections should be examined (blind) by another pathologist to confirm the diagnoses.
See USEPA (1987b) for additional information on QA/QC procedures for
histopathological examinations.
4.6.5 Statistical Design Considerations
Statistical strategies may mitigate the high costs of pathobiological monitoring methods.
As discussed earlier, power analyses considering the strategy of compositing samples
can often lead to a cost-effective monitoring design strategy. Power-cost analyses are
necessary in selecting the appropriate sample/replicate number, sample location, and
sampling frequency. If the primary objective of a monitoring program is to determine
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pathobiological differences between sampling locations, composite sampling may be an
appropriate strategy. However, a limitation of composite sampling is the inability to
directly estimate the range and variance of the population of individual samples.
Similarly, the use of space- or time-bulking strategies will severely limit a monitoring
program's ability to assess spatial and temporal heterogeneity of the samples.
Given that the monitoring program must accommodate a fixed level of sampling cost,
the best strategy for pathobiological monitoring is to collect more replicates at fewer
locations. For histopathology, from 25 to 100 animals per species may need to be
surveyed to detect low incidences of disease. The selection of appropriate statistical
analyses must be made with these limitations in mind, as well as the specific types of
statistical analyses that can be performed on different types of data. For example,
quantitative information from measurements of enzyme levels or computerized image
analysis/morphometric programs may be effectively analyzed by parametric statistics (if
the data fit the conditions for normal distributions, limitations of assumptions, etc.), but
histopathological observations may need to be rated on presence/absence or
categorized qualitative scales that will require nonparametric techniques. USEPA
(1987b) provides additional information on the types of sampling designs and statistical
tests that may be appropriate for these data.
4.6.6
Use of Data
Data Interpretation
Data interpretation for pathobiological methods may be limited because there are
relatively few trained personnel, facilities and equipment may be expensive, and
references for newly emerging techniques may be scarce. However, because of the
recent interest in developing biomarkers for monitoring the effects of natural and
anthropogenic environmental stresses on aquatic organisms, many research programs
are under way at Federal, State, and academic facilities to develop, standardize, and
validate the most promising biomarkers for sentinel species that will establish
cause-and-effect links for pollutant exposure. Basic laboratory research and
experimental studies must be conducted in conjunction with field work to elucidate the
relationships between contaminant levels, structural, biochemical, or functional
pathologies, and population health (Johnson and Bergman, 1984; McCarthy, 1990). A
critical concern is the coordination of efforts and creation of a multidisciplinary approach.
It is important to establish baseline data on selected species, as well as to demonstrate
the alterations in the health of those species due to contaminant exposures. Methods
and techniques, terminology, and interpretation of lesions and effects must be
standardized. In addition to a rapidly expanding body of literature on baseline measures
of health, histological atlases for several species of fish and shellfish, and courses,
workshops, meetings, and special symposia (e.g., Responses of Marine Organisms to
Pollutants/Woods Hole, MA, and Plymouth, England; Annual Aquatic Toxicology
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Workshop, Canada), several professional societies (e.g., Society for Invertebrate
Pathology, American Fisheries Society/Fish Health Section) are facilitating training in
techniques and communication among investigators and laboratories. Diseases of fish
and shellfish have been reviewed by several authors (e.g., Sparks, 1985; Ferguson,
1989; Roberts, 1989; Sindermann, 1990; Couch and Fournie, 1993). For
histopathological studies of tumors in fish, standardization of nomenclature will be
available in Dawe et al. (in press).
As for any monitoring program, long-term studies of health and disease in aquatic
organisms will aid in identifying and interpreting observed pathobiological effects.
Furthermore, it will be important to archive data so that they will be available for future
comparisons. The Registry of Tumors in Lower Animals at the National Museum of
Natural History, Smithsonian Institution, Washington, DC, houses over 3,500 cases of
neoplastic and nonneoplastic lesions representing a wide variety of host aquatic
species, pathogens, and environmental stresses from field and laboratory studies
conducted around the world. The Registry also contains the Registry of Marine
Pathology, originally developed by the National Marine Fisheries Service (NMFS), and
the collection of crustacean histopathological material researched by Dr. Phyllis T.
Johnson, NMFS. Although primarily serving as a clearinghouse for information on
neoplastic diseases and research, the Registry is also able to direct inquiries for
information on nonneoplastic diseases, and the materials archived there are available
for study by qualified investigators. Data collected from the long-term monitoring efforts
of NOAA's National Status and Trends Program (NS&T) and EPA's Environmental
Monitoring and Assessment Program (EMAP) will also provide useful information for the
interpretation of the various pathobiological methods that are being employed in these
studies.
Data Integration
Measuring or evaluating the effects of stressors on fish and shellfish by pathobiological
indicators is difficult primarily because of the large number of variables that can
influence biological response. Variables including water temperature, nutritional status,
species, sex, reproductive and developmental stages, and physiological functions can
render tests difficult to compare and evaluate.
Anderson (1990) outlined a plan including four levels of study in a tiered approach to
investigate the effects of stress on the immune system and to quantify the possible
contribution of these environmental and physiological variables on the stress response.
The plan, generally speaking, appears to be applicable to all pathobiological methods
and may be adapted to them using the following four levels: (1) observations of fish and
shellfish populations in the field; (2) studies of caged fish or shellfish in the field; (3) in
vivo exposures in the laboratory; and (4) in vitro assays in the laboratory.
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The first level of study includes the collection of information from field observations on
relative abundance, reduction in sport-fish catches, or decline in commercial harvests.
Reduced yields reported by anglers in previously productive fishing grounds may be
correlated with environmental stressors or pollutants occurring at these sites. If sick or
moribund fish or shellfish are called to the attention of local biologists by anglers, gross
morphological descriptions might be made and isolation of the disease agent attempted.
The second level of investigation involves the use of caged fish or shellfish placed in the
field to test for the presence or action of environmental stressors. Groups of fish or
shellfish can be placed in pens or cages at suspected sites and their responses (signs of
disease, acute mortality) can be compared with those of control fish or shellfish caged in
unstressful areas. Organisms from the cages can be sampled at various intervals and
the intensity of various immune responses quantified as a function of the time the fish or
shellfish were exposed. Levels of immune responses can be compared by injecting
specific disease agents into both control and test organisms, and recording disease and
death rates.
The third level of assay is in vivo laboratory tests, by which immune response can be
evaluated in experiments with calibrated dilutions of specific contaminants and kinetic
measurements of the immune response at each dilution. Challenges with disease
agents can be more easily controlled in the laboratory to provide information on how the
stressor makes the fish or shellfish more susceptible to specific pathogens.
A fourth level of investigation is the recently developed method for testing the effects of
chemicals, drugs, and other stressors in vitro. Spleens and other immunopoietic organs
can be removed from fish and placed in tissue culture media and their reactions to
pollutants tested using the passive hemolytic plaque assay (Anderson et al., 1986). Use
of this method to measure the effects of stressors allows maximal control of the levels of
pollutants. Because the immune response is monitored under defined laboratory and
environmental conditions, important information is obtained about how specific stressors
affect the immune response. In vitro methods require fewer fish than in vivo methods
because tissue and organ samples can be divided into many sections, which also
reduces the variability of responses.
McCarthy (1990) presented a research strategy to validate biomarkers and provide the
scientific understanding necessary to interpret biomarker responses. An evolving
monitoring program was proposed that focused broadly on evaluation of contamination
in an array of ecosystem types. The challenges and obstacles to be addressed in such
a program include the following:
• The quantitative and qualitative relationships between chemical exposure,
biomarker response, and adverse effects must be established.
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• Responses due to chemical exposure must be distinguishable from natural
sources of variability (e.g., ecological and physiological variables,
species-specific differences, and individual variability) if biomarkers are to be
useful in evaluating contamination.
• The validity of extrapolating between biomarker responses measured in
individual organisms and some higher-level effect at a population or
community level must be established.
• The use of exposure biomarkers in animal surrogates to evaluate the
potential for human exposure should be explored.
The plan is an ambitious multiyear research and development program involving the
formulation of a long-term, interagency, interdisciplinary activity such as the
Environmental Monitoring and Assessment Program (EMAP) by EPA in cooperation with
other Federal agencies.
4.6.7 Summary and Recommendations
Rationale
• Pathobiological methods provide information concerning biological organ
systems that, through an evaluation of organ structure, activity, and function,
can be used to determine adverse effects of pollutants in the environment.
• Pathobiological methods should be used in concert so that cause-and-effect
relationships can be evaluated.
Monitoring Design Considerations
• Sampling stations should be located along a contamination gradient (i.e.,
from highly contaminated to uncontaminated). This type of sampling strategy
will allow dose-response relationships to be evaluated.
• Fish pathobiological monitoring should be conducted only if the target
species could reasonably be expected to have spent a considerable amount
of time within the area of contamination.
• Large sample sizes will frequently be required for fish pathobiological
monitoring due to natural variability among individuals and taxa.
Existing Analytical Methods
Anatomic Pathology Methods
• Tissues are examined with the naked eye or the aid of a microscope
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• Light microscopic (LM) methods
- detect changes at the cellular level
- organ and time-series specific
• Electron microscopic (EM) methods
- detect subcellular changes
- specific as to area of cell or site of pollutant action
• Histochemical methods
- detect enzyme systems or reaction products in the cell
- specific to chemical class
Reproduction/Development Methods
• Reproductive capacity is reflected in population recruitment and abundance
. Methods that examine the egg itself provide a direct measure of reproductive
effects from pollutants
• Cytogenetic (DNA) tests
- in vitro tests offer greatest flexibility in terms of sensitivity, expense, and
time required and are frequently conducted for convenience and
availability.
- sister-chromatid exchange (SCE) assays are DNA damage tests used as
a screening tool and are dose-responsive and sensitive to low
concentrations of pollutants.
- the aneuploidy technique is simple and easy to use but is inaccurate for
sublethal effects.
Biochemical Methods
• Several methods are specific for a certain compound class.
• Use a suite of indicators for environmental monitoring.
• Microsomal mono-oxygenase (MO) assays could serve as sensitive
biological indicators for certain chemical classes.
• Metallothionein (MT) assays are nonspecific for metal exposure, but can
provide information on the likelihood of a particular metal pool producing a
pathological effect.
• Stress proteins are not pollutant-specific, but could result from generalized
stress.
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Immunologic Methods
• Immunologic assays can provide information on pollutant-induced stress
effects; however, they may require confirmatory assays.
• Many animals should be sampled because of natural variability among
animals.
Physiologic Methods
• Serious "interferences" can be caused by stress induced during collection
and may limit the potential usefulness of physiological tests because effects
of toxicants cannot be distinguished from those induced during handling of
the wild organisms.
• Hematologic methods for fish, such as measurements of hematocrit (packed
cell volume), hemoglobin concentration, erythrocyte count, and leucocyte
count (or volume), can be successfully used. Although the fish are not
immune to stress of capture, hematologic methods are influenced far less
than some other measurements such as blood glucose.
• The leucocrit can be a more sensitive measure of metal pollution than the
hematocrit.
QA/QC Considerations
• Pathobiological methods have a wide range of sensitivities.
• Organisms must be carefully handled and properly prepared for each method.
• The main areas of concern with regard to analytical QA/QC are precision,
accuracy, representativeness, completeness, and comparability.
Statistical Design Considerations
• Composite tissue sampling consists of mixing tissue samples from two or
more individuals collected at a particular location and time.
• Space-bulking (combining composites from several locations) and/or
time-bulking (combining several composites over time from one location)
strategies should be used judiciously since information concerning spatial
and temporal heterogeneity may be lost.
• Power analyses have shown that for a fixed level of sampling effort, a
monitoring program's power is generally increased by collecting more
replicates at fewer locations.
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The appropriate statistical tests must be performed and may vary depending
on the type of data generated.
Use of Data
Data use and interpretation may be limited by relatively few trained personnel
and expensive equipment, but many research and training programs are now
under way at Federal, State, and academic facilities to improve biomarker
methods in sentinel species.
Basic laboratory research must be conducted and biological methods must
be tested in the field.
Information communication
- need to coordinate efforts, communicate, and create a multidisciplinary
approach to relate lab and field studies and establish cause-and-effect
relationships.
Data integration
- long-range research strategies should be followed to validate biomarkers
and provide scientific understanding necessary to interpret biomarker
responses.
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4.7
FISH POPULATIONS
The objective of monitoring fish populations is to detect and describe spatial and
temporal changes in the abundance, structure, and function of fish communities. To
attempt to protect and preserve healthy fish from the possible effects of pollutant
discharges, estimates of fish population abundance and detailed knowledge of fish life
histories are required.
Except in the most severe cases, it is usually impossible to directly link pollutant
loadings to stock levels (i.e., there is a poorly established cause-and-effect relationship
between pollution and fish population responses). Natural population fluctuations in
fishery stocks will usually be much greater than effects due to pollution. The effects of
discharges are obvious when fish are killed outright by high concentrations of toxic
substances (Rothschild, 1986). However, these instances are rare compared to the
prevalent situation, in which fish are exposed to lower pollutant levels that may not be
lethal but are thought to be harmful in some sense. Sublethal pollutant loadings can be
expressed in terms of increased body burden of toxicants or greater incidence of various
lesions. The effects of sublethal concentrations on growth, mortality, and reproduction
are not known (Rothschild, 1986) As a result of these limitations, fish population
measurement methodologies require further refinement and field validation before they
can be promoted for regular use in 403 monitoring or permit decisions.
4.7.1
Rationale
Under section 403 of the Clean Water Act, permitters must use 10 guidelines in
determining whether a discharge results in unreasonable degradation or irreparable
harm to the marine environment. Fish population studies have generally not been used
to directly link a source of pollution to observed responses. However, some recent
studies have indicated that selection of resident bottom fish as indicators of pollution can
support other data linking sources of pollution to environmental effects, especially if such
studies are conducted in conjunction with benthic studies and parallels can be drawn
between the two. In addition, such methodologies do address the following three 403
guidelines:
• Composition and vulnerability of potentially exposed biological communities;
• Importance of the receiving water area to the surrounding biological
community; and
• Existing or potential recreational and commercial fishing.
Population effects can include those caused by changed reproductive rates or changed
distribution and migration patterns. Effects of suppressed reproductive rates on
population density have been clearly demonstrated in the diminishing populations of the
striped bass in San Francisco Bay. The suppressed reproductive rates have been
shown to correlate with toxic concentrations in the bay (Whipple et al., 1984). Several
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laboratory studies have demonstrated that certain fish can avoid toxics, including DDT,
endrin, and Duroban. There is, however, no proof that fish avoid toxicants in the field. If
they do, "hotspots" would be avoided by such species and the population would suffer
less exposure to toxic effects.
One of the greatest problems in analyzing effects of toxics on populations is separating
direct toxic effects from natural or non-pollution-related variations including climatic
changes and overfishing. Results of population studies must be evaluated together with
measurements of other environmental parameters before any cause-and-effect
relationship can be established.
4.7.2 Monitoring Design Considerations
Specimen collection, analysis, and evaluation of fish community structure and function
are typically time-consuming, labor-intensive, and expensive tasks. A survey vessel
manned by an experienced crew and specially equipped with gear to collect organisms
is required. Expert taxonomists are needed to identify and enumerate collected fish
specimens.
The results of fish monitoring programs can vary substantially depending on the
objectives and corresponding design specifications. The characteristics primarily
responsible for the variability in the results are the following:
• Type of sampling gear,
• Volume sampled, and
• Location and timing of sample collection.
It is essential to understand the effects of these monitoring design characteristics on the
results and to standardize them as much as possible to ensure the comparability of
sampled data. Analyses of power-cost efficiencies are useful in selecting the
appropriate sampling gear and sample processing protocols. Ferraro et al. (1989)
provide an example of power-cost analyses.
Sampling Devices
Sample collection protocols influence all subsequent laboratory and data analysis; it is
key that fish population samples be collected using acceptable and standardized
techniques. Several types of devices can be used to collect fish samples: traps and
cages, passive nets, trawls (active nets), and photographic surveys (Fredette et al.,
1989). Many of these devices selectively sample specific types of fish. Accordingly,
conducting comparisons between data collected using different devices and even
different nets of the same type is inadvisable.
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Traps and Cages
Traps and cages are usually designed to attract and capture specific organisms. They
are useful in studies examining the activities of a particular target organism in a given
area. Traps and cages provide only qualitative measures of organisms in a particular
area.
Passive Nets and Trawls
Nets vary in their selectivity of the species that are captured and in the efficiency of
retaining captured specimens. The size of the net, its configuration and orientation, and
the avoidance behavior of the target species should be considered when using any net.
The mesh size affects the speed at which the net can be towed, as well as the size of
fish caught. The slower the towing speed, the more likely that some organisms will
either avoid or escape the net. It is highly recommended that the mesh size, net
opening, and duration, direction, and speed of towing be set in order to compare trawls.
In any monitoring program, otter trawl net size must be kept constant to ensure
intercomparison of sampling results.
Passive nets (e.g., gill nets) are deployed at a fixed position; organisms become
entangled or trapped within the netted area. Passive nets are used to collect selected
target species and to provide a qualitative means of sampling fish populations.
Limitations associated with passive nets include: .
• Nets, ordinarily, must remain in place for an extended period of time.
• Deployment and recovery of nets are typically time-consuming processes.
Trawls (active nets) are drawn through the water, and results are more immediate than
those obtained through the use of passive nets. Trawls are typically used to collect
large quantities of fish at various depths (Fredette et al., 1989).
Photographic Surveys
Photographic surveys are effective when the bottom topography is uneven or trawling is
not possible. The utility of photographic surveys is limited by water clarity, difficulties in
identifying species, and fish avoidance of the camera system. In addition, further
studies comparing photographic surveys to trawls are required before comparisons may
be made. This method is also limited by the qualitative nature of the data (Fredette et
al.,1989).
Volume Sampled
Different species of fish have different scales of horizontal and vertical spatial
distribution (Gushing, 1975; Bond, 1979). Costs of laboratory analysis of the sample
increase with increased volume sampled. Analyses of spatial and temporal scale,
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statistical power, and costs will assist in determining optimal sample volume. It is highly
recommended that a standard sample volume (same tow duration, tow speed, and net
opening area) be analyzed to ensure data comparability (Green, 1979).
Selection of Sampling Period
Fish assemblages are dynamic; the most common temporal patterns observed in fish
communities are those associated with daily activity patterns (diurnal), seasonal
changes, and life history strategies. To minimize energy expenditure, most fishes mimic
the diurnal activities of their prey. These diurnal fluctuations typically occur in the
vertical scale and should be considered to ensure collection of representative samples.
Seasonal variation in fish assemblages may be due to changes in physical, chemical,
and/or biological parameters: i.e., temperature, light transmissivity, dissolved oxygen,
predation, reproductive stage, recruitment, etc.
Given the seasonal variation characteristic of fish assemblages in general, it is
recommended that direct comparisons between samples collected during different
seasons be avoided. Studies investigating interannual variation in the characteristics of
fish communities should be conducted during the same season (preferably the same
month) each year.
4.7.3 Analytical Methods Considerations
There are a variety of approaches to assess the effects of anthropogenic perturbance on
fish communities of the marine environment. These assessment approaches may be
grouped into three categories :
• Biological indices,
• Indicator species, and
• Multivariate analyses.
There has been little consensus among biologists regarding the suitability of various
techniques for describing community characteristics and/or for assessing impacts. A
critical evaluation of the use of biological indices to detect environmental change is
presented in an EPA Technical Support Document (USEPA, 1985d). The indices,
shown in Table 4-22, are evaluated on the basis of the following criteria :
• Biological meaning,
• Ease of interpretation, and
• Sensitivity to community changes due to anthropogenic sources.
The results of these evaluations and additional information on other analytical methods
are summarized below.
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Table 4-22. Biological Indices
Index/Method
Biological Characteristic
Measured
Bray-Curtis (Curtis and Peterson, 1978)
Dominance (Swartz et al.,1985)
Number of individuals (USEPA,1985d)
Number of species (USEPA.1985d)
Biomass (USEPA,1985d)
Margalefs SR (Margalef,1969)
J (Pielou,1966)
Shannon-Wiener H' (USEPA,1985d)
Dissimilarity
Community structure
Total abundance
Total taxa
Standing crop
Diversity
Evenness
Diversity
Biological Indices
The number of individuals and the number of species have been used as indicators of
anthropogenic disturbance, as well as other environmental stresses (USEPA, 1985d).
Furthermore, these simple biological indices are less ambiguous and can often be as
informative as diversity indices (USEPA, 1985d; Green, 1979; Hurlbert, 1971).
Measures of biomass have inherent problems in the collection of the data, e.g., loss or
gain of weight due to preservative medium, drying times, or evaporative weight loss.
More complicated indices-e.g., species diversity, species richness, dominance,
evenness-have found varying degrees of acceptance. Because of this, these methods
should be used in conjunction with other assessment methods to help verify results.
Diversity indices, which are measures of the distribution of individuals among species,
have the following limitations (Green, 1984b):
• They often lack biological meaning;
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• They are not robust empirical indicators of any important correlates of
environmental health;
• They do not incorporate information on form and function of resident species; and
• They are susceptible to biases associated with well-described taxa.
Even so, species diversity indices are a widely used measure of community structure.
Although diversity may not be able to pinpoint a single problematic discharge in an area
that contains many, it can be an indicator of degraded water quality. Species diversity in
fish communities is noted to increase with increasing distance from a pollution source
(Armstrong et al., 1970; Bechtel and Copeland, 1970; Tsai, 1968). In fact, Bechtel and
Copeland (1970) found a linear relationship between the volume of toxic effluent
entering the Houston Ship Channel and the fish diversity of shallow stations.
The dominance index is a measure of the degree to which one or a few species
dominate the community. The dominance index, herein defined as the minimum
number of species required to account for 75 percent of the total number of individuals,
has been useful in describing community structure (Swartz et al., 1985). Its advantages
are that it is easily calculated, it does not assume an underlying distribution of
individuals among species, and it is statistically testable.
Indicator Species
The evaluation of abundances of individual indicator species is generally informative and
may reduce the cost of the analysis. The absence of pollution-sensitive species and the
enhancement of opportunistic and pollution-tolerant species may assist in defining the
spatial and temporal extent and magnitude of impacts. A preponderance of
unspecialized feeders indicates areas of stress. However, indicator variables must
possess the following characteristics (Green, 1984b):
• They must provide a sufficiently precise and accurate appraisal of:
- species of concern,
- anthropogenic disturbances to communities, and
- presence/absence or magnitude of anthropogenic perturbance to the
ecosystem.
• They must provide a cost-effective and statistically reliable alternative to
monitoring all critical measures of habitat perturbance.
• They must be appropriate for the spatial and temporal scale demanded by
the study objectives.
Further studies of the response patterns of fish species subjected to anthropogenic
perturbations caused by discharges are required in order to select appropriate indicators
of environmental impact. If a suitable indicator species is identified, it is desirable to
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monitor the status and trend of that species' population. This will also be true of certain
species having high economic or public value (e.g., striped bass). A wide variety of
methods can be used to measure the size of fish populations and assess population
structure. These include mark and recapture techniques and various techniques used to
determine population sex and age structure. Nielson and Johnson (1984) and Cailliet et
al. (1986) provide thorough discussions of these techniques and their applications.
Bicker (1975) provides an excellent discussion of methods for the sampling of fish
populations.
Statistical Analyses
Numerical classification encompasses a wide variety of techniques that have been used
in the analysis of fish data to distinguish groups of entities (e.g., sample locations)
according to similarity of attributes (e.g., species). These techniques differ from most
multivariate methods in that no assumptions are made concerning the underlying
distributions of the variables. Detailed descriptions of numerical classification analysis
can be found in Romesburg (1984), Clifford and Stephenson (1975), Boesch (1977),
Sneath and Sokal (1973), arid Anderberg (1973). Boesch (1977) is particularly valuable
as an introduction and guide to the use of numerical classification analysis in marine
environmental studies. Guidance on the interpretation of classification results is
provided in an EPA Technical Support Document (USEPA, 1988d).
Ordination analyses have also been used to reduce the dimensionality of the data set
while maintaining the relationship between similar and dissimilar entities. At present no
single ordination technique has been shown to be clearly superior for the analysis of
biological data (USEPA, 1985d).
Multivariate analyses are effective heuristic tools. They generate visual representations
that often indicate where further analyses ought to be conducted.
Analytical Approach Recommendations
Some of the most informative measures of community structure are the simplest (Table
4-22):
• Number of individuals,
• Number of species,
• Dominance,
• Abundance of pollution-sensitive species, and
• Abundance of opportunistic and pollution-tolerant species.
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These indices have proved to be useful over various habitats and regions in assessing
changes in fish population dynamics (USEPA, 1985d). Values of these variables may
be determined from the list of species abundances generated during the taxonomic
identification of collected specimens. Furthermore, the values of these variables may be
easily tested statistically using parametric or nonparametric techniques. It is
recommended that no single index or analytical method be used to assess impacts;
rather, the assessment of impacts should incorporate information that each variable and
method contributes concerning community structure.
Selection of reference sites is key to the evaluation of environmental impact assessment
of point source dischargers. Results of analyses using reference measures provide the
means of comparison by which anthropogenic impacts are detected. It is essential that
selected reference sites exhibit similar habitat characteristics compared to monitoring
program sampling sites. Several reference sites may be required to meet these criteria.
4.7.4 QA/QC Considerations
Sample Collection
Nets should be inspected for wear and tear leading to sample loss.
Taxonomic Identification
A key QA/QC issue is taxonomic standardization. Consistent taxonomic identifications
are achieved through interaction among taxonomists working on each major group.
Participation of the laboratory staff in regional taxonomic standardization programs is
recommended to ensure regional consistency and accuracy of identification.
A sampling of the verified specimens should then be placed in a permanent reference
collection. All specimens in the reference collection should be stored in labeled
containers that are segregated by species and sample. Reference specimens should
be archived alphabetically within major taxonomic groups.
4.7.5 Statistical Design Considerations
Consideration of statistical strategies will mitigate the high costs of collecting and
processing samples. Power-cost analyses are necessary in selecting appropriate
sample/replicate number, sample location, and sampling frequency.
Temporal Stratification of the Data
The time of the year should be controlled or stratified in the design; the use of annual
averages is seldom good practice. Temporal stratification of the data should not be
attempted until sufficient knowledge of long-term natural cycles is attained. Initially,
simple regression analyses may be conducted on seasonally stratified data to identify
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monotonic temporal trends. Whether conditions are improving or degrading over time
may be further examined using various statistical time series analyses (e.g., temporal
autocorrelation, spectral analyses).
Statistical Power
Selection of the number of replicates is an important component of program design.
The inherent patchy distribution of fish communities requires collection of sufficient
replicate samples to ensure an accurate description of the area. However, increases in
replication increase sample processing costs. Power analyses assist in the allocation of
sampling resources (stations, replication, and frequency) with regard to program
finances and design.
Power analyses may be applied to determine the appropriate number of sample
replicates required to detect a specified difference (USEPA, 1987d). The number of
replications required to detect a specified minimum difference is a function of the
statistical power and the variance in the data. Power analyses require a priori
knowledge of the variability in the data. A best guess or, preferably, variation observed
in historical data is often used initially in the design of the monitoring program.
To improve the power of a statistical test, while keeping the significance level constant,
the sample size should be increased. Because of constraints in cost and time imposed
by the monitoring program, however, this option may not be available. Power analyses
have shown that for a fixed level of sampling effort, a monitoring program's power is
generally increased by collecting more replicates at fewer locations. The number and
distribution of sampling locations required to evaluate the effectiveness of the monitoring
program will depend on the size and complexity of the discharge and the surrounding
environment.
4.7.6
Use of Data
Monitoring of fish community structure provides a measure of the health of the marine
habitat and can be used as a tool in the evaluation of spatial and temporal effects of
marine discharges. The presence or absence of certain fish is useful in indicating the
condition of the environment. Monitoring of fish communities also provides data
required in the design and validation of fish population dynamics models and the
selection of biological indicators (USEPA, 1990d).
In addition, monitoring of fish communities may directly provide accurate information
essential in assessing the effectiveness of pollution abatement programs. Analyses of
fish communities may be effectively used to monitor long-term change in the receiving
environment (Gushing, 1975).
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Fish serve as effective indicators of overall environmental condition, delineating the
magnitude, spatial extent, and temporal trends of anthropogenic and natural
perturbations to the ecosystem. Monitoring of fish will provide relevant accurate data
fundamental to achieving the objectives of many marine monitoring programs. Such
measurements, however, will not provide information linking specific pollution sources to
observed effects. The results of the fish stock monitoring program can also be unclear
as a result of the mobility of fish and their range of natural fluctuation. The data can be
used to augment other monitoring data to obtain a clearer picture of the discharge
effects.
4.7.7 Summary and Recommendations
Rationale
• The objective is to detect and describe spatial and temporal changes in the
size, structure, and function of fish communities to protect the economic,
recreational, and aesthetic value of the resident fisheries.
• Monitoring provides in situ measures of habitat quality and is a powerful tool
in assessing environmental impact.
Monitoring Design Considerations
• It is recommended that the types of sampling gear, the volume sampled, and
the location and timing of sample collection be consistent to allow for
comparisons among studies.
• To reduce the variation due to seasonal differences, sampling should be
conducted during the same season—preferably the same month—each year.
Analytical Methods Considerations
• It is recommended that simple measures of community structure be used to
assess the condition of the marine fish: number of individuals, number of
species, dominance, abundance of pollution-sensitive species, and
abundance of pollution-tolerant species.
• Indicator species should possess the following characteristics:
- sensitive to perturbances of concern;
- cost-effective and statistically reliable alternative to measuring all species
in a monitoring program;
- statistically reliable indicative measures of habitat perturbance; and
- appropriate for the spatial and temporal scale demanded by the study
objectives.
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• Selection of reference sites is key to the evaluation of environmental impact
due to discharges; several reference sites may be required to provide proper
control for sampling sites.
QA/QC Considerations
• Taxonomic standardization is key to the analysis of community structure.
Recommended protocols include consistent interactions among taxonomists,
reidentification of selected samples, and use of a reference collection.
Statistical Design Considerations
• Power analyses may be applied to determine the appropriate number of
sample replicates required to detect a specified difference, thereby optimizing
the high costs of collecting and processing samples.
Use of Data
• Data provide essential information to assess impacts due to anthropogenic
perturbance, monitor recovery of the receiving environment, and validate
community and population models.
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4.8
PLANKTON: BIOMASS, PRODUCTIVITY, AND COMMUNITY
STRUCTURE/FUNCTION
Although increased primary production resulting from intentional nutrient inputs has
been shown to increase fish stocks in some experimental systems, the possible
increase in fisheries in naturally productive systems is considered insignificant. In fact, it
is more likely that stocks will decrease due to modification of the plankton species
composition. Biogeographical distributions of plankton are determined by specific
environmental factors such as light, temperature, salinity, and nutrients. Plankton vary
in size over a range of seven orders of magnitude. Each predator species can utilize
only plankton within two or three orders of magnitude in size, at the most, as a food
resource. Additionally, the nutrient values of various prey differ so that a change in
species composition to less nutritive species may cause stress on the predator species.
The purposes of monitoring plankton characteristics are (1) to identify the dominant species;
(2) to detect short- and long-term spatial and temporal trends in overall biomass and
productivity and species abundance, distribution, and composition; and (3) to examine the
relationship between water quality conditions and these characteristics.
As was the case for fish population monitoring methods, measurements of plankton
biornass, productivity, and community structure/function are not sufficiently developed to
be used in a discharge-specific monitoring framework for section 403 assessments at
this time. These measurements suffer from a lack of a specific cause-effect relationship
between pollution and plankton responses.
4.8.1
Rationale
Although not directly applicable for use in monitoring the site-specific impacts of
individual point source discharges, plankton measurements can be used to evaluate 3 of
the 10 guidelines presented in the section 403 regulations for assessing the potential for
unreasonable degradation:
• Composition and vulnerability of potentially exposed biological communities;
• Existing or potential recreational and commercial fishing; and
• Importance of the receiving water area to the surrounding biological
community.
Changes in nutrient concentrations in coastal waters may result in the potential for
long-term biological changes in the plankton community that may lead to changes in
species distribution and abundance (both primary producers and consumers). A
concern of the section 403 program is that excessive nutrient inputs can result in
excessive phytoplankton biomass, which in turn can lead to increased turbidity and/or
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changes in phytoplankton species composition and perhaps the trophic structure. If,
based on water quality monitoring data, changes in trophic structure are suspected of
occurring, periodic monitoring of the plankton community may be required.
The plankton component of a monitoring program could be used to provide data necessary
to assess the effectiveness of management programs in mitigating potential adverse effects
due to changes in the plankton community biomass and structure/function.
4.8.2 Monitoring Design Considerations
Plankton
Plankton monitoring strategies should be able to delineate between natural or seasonal
variability in plankton stocks and variations caused by changes in nutrient concentrations.
Characterization of phytoplankton taxonomic abundance and distribution and primary
productivity provides indications of water quality conditions. Monitoring changes in
phytoplankton population composition and densities is critical for the interpretation and
evaluation of long-term trends in water and habitat quality. Further understanding of the
causes of excessive water column and sediment oxygen demand requires tracking of
photosynthetic activity and metabolic rates over time (Chesapeake Executive Council,
1988b). Zooplankton abundance and distribution are affected by both changes in
phytoplankton and changes in predator populations. Therefore, population characteristics of
this group can indicate symptoms of water quality problems, fishing pressure, and other
habitat problems for predator species (Chesapeake Executive Council, 1988b).
Selection of Temporal Sampling Strategies
Because of their short turnover times, phytoplankton communities may respond to
perturbations much more rapidly than other biotic groups. Therefore, phytoplankton
samples should be collected relatively more frequently. In those situations where
phytoplankton communities display pronounced seasonal variations in standing stock or
production, it may be appropriate to use a temporally stratified sampling approach. For
example, in the Maryland Chesapeake Bay Phytoplankton Monitoring Program, sampling
takes place once monthly from October through March and twice a month from April through
September (USEPA, 1989c).
The life span of zooplankton, on the other hand, is longer than that of phytoplankton, so the
capacity for responding to perturbations is less than that of phytoplankton. Therefore, less
frequent sampling is required. As with the phytoplankton community, the zooplankton
monitoring program should consider the natural temporal fluctuations in abundance and
species composition.
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In many cases, regular monitoring of the zoopiankton community may not be necessary
unless changes in the phytoplankton community that would induce changes in the herbivore
community are observed. Because many zoopiankton graze on phytoplankton, in areas
where the phytoplankton community has been affected, alterations of the zoopiankton
community are a distinct possibility. In other cases, monitoring of zoopiankton may be
desirable only when there is evidence of previous impact on the zoopiankton community or
in those situations where point and nonpoint source discharges are located in areas where
there is high potential impact on zoopiankton (for example, in environments with
macroplanktonic larvae of important commercial or recreational species).
Sampling Methods
Phytoplankton
Phytoplankton samples should be collected at a variety of depths throughout the water
column, including some above and some below the pycnocline. For example,
composite water column samples collected for the Virginia Chesapeake Bay Plankton
Monitoring Program are taken from five depths above and five depths below the
pycnocline (USEPA, 1989c). Vertical collection of chlorophyll a (an indirect measure of
phytoplankton standing stocks) can be examined at such stations through the collection
of water samples with water bottles at various depths followed by fluorometric or
spectrophotometric determination of chlorophyll a. If available, a pump station may be
used with a flow-through fluorometer for a continuous profile of chlorophyll a
concentration with depth (Lorenzen, 1966). Samples to be used for taxonomy analysis
should be collected with water bottles because agitation associated with pumping may
damage cells, making them unidentifiable. Pumps should be used only for
determination of chlorophyll a concentrations.
Continuous vertical profiles of chlorophyll a concentrations can be obtained by attaching
a fluorometer to a conductivity-temperature-depth (CTD) system. An additional sensor
measuring dissolved oxygen can be attached to the CTD system. This combined
instrument package will make direct measurements of chlorophyll a levels and water
column stratification. The advantage of this measurement system is that instead of
pulling water up from depth to a measuring device on the deck of a ship, the instrument
is lowered through the water column, taking samples in undisturbed water. Sampling
rates can be as high as 24 samples per seconds with a locating speed of 1 meter per
second. Data are usually collected by a computer-based data acquisition system; in
advanced systems, data can be displayed in real time as they are being collected. A
rosette water sampler can be lowered at the same time as the CTD package to collect
water samples at discrete depths. These samples will be used for taxonomy analysis
and to verify and calibrate chlorophyll a measurements from the CTD fluorometer. The
calibration will be carried out using laboratory spectrophotometer methods.
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Zooplankton
Because zooplankton possess varying degrees of swimming ability, they have the
potential for aggregating in patches or in a narrow depth strata. This introduces
additional complications into quantitative sampling. It also means that zooplankton are
able to avoid certain types of gear. Until recently, the role of the microbial loop, which
consists of bacteria, flagellates, ciliates, and microzooplankton (<200 |im), has been
overlooked. These organisms may represent a significant pathway in the reutilization
and conversation of dissolved organic carbon into larger zooplankton and benthic
organisms (Azam et al., 1983; Pomeroy, 1984). The sampling methods to be used for
collecting zooplankton will vary depending on the size of the organisms.
Microzooplankton (size range of 20 - 200 ja,m) can be collected with water bottles at
various depths similar to those used for phytoplankton (Jacobs and Grant, 1978), or
small (for example 44 |o,m) rnesh nets can be used. Pumping systems can also be used
(Beers et al., 1967). These have the advantage of being able to take samples while the
ship is under way, but they may damage soft-bodied organisms and they are more
expensive and complicated than water bottles. Triplicate samples should be collected
from each station depth, thus allowing for statistical analysis of intrastation variability.
For small mesozooplankton (greater than 200 |im), nets are usually used (UNESCO,
1968). Additional tows may have to be made with larger nets in order to collect
representative samples of larger zooplankton and larval fish. All tows should be
replicated. The number of replicates necessary for the desired precision of estimation
should be determined during a preliminary or pilot sampling program. A number of other
considerations, including net mouth diameter, towing speed, and shipboard handling of
samples, will affect sampling results. Some problems associated with the use of nets for
zooplankton sampling include avoidance, which may result in underestimating
abundance and diversity (McGowan and Fraundorf, 1966; Wiebe and Holland, 1968),
and clogging and loss of filtration efficiency.
4.8.3 Analytical Methods Considerations
Phytoplankton
Biomass and Productivity
Phytoplankton biomass can be indirectly measured through the measurement of the
concentration of chlorophyll a in the water. This is done through fluorometric or
spectrophotometric measurements. Within these methods are differences in extraction
techniques for chlorophyll determination, including various methods of filtration,
solvents, temperature, and/or physical treatment (sonication or grinding) (D'Elia et al.,
1986).
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The use of chlorophyll a measurements, especially fluorometric, has become
widespread primarily because the method is relatively fast, simple, and reproducible.
With both the fluorometric and spectrophotometric determinations, however, the
presence of accessory pigments may interfere with determination of chlorophyll a.
Several researchers, nevertheless, have successfully used chromatographic procedures
to separate interfering substances prior to determination (D'Elia et al., 1986). High
Performance Liquid Chromatography (HPLC) is generally acknowledged as the most
accurate (and most expensive) method for chlorophyll determinations (C.F.
Zimmermann, 1990, Chesapeake Biological Laboratory, Solomons, MD, personal
communication).
Other methods for estimating phytoplankton biomass used in earlier monitoring surveys
include cell counts, total cell volume estimates, protein estimates, and dry weight.
These and other methods have certain disadvantages related to speed of the technique
or degree of accuracy (D'Elia et al., 1986).
Phytoplankton primary productivity should be measured by the 14C light-dark bottle
technique (UNESCO, 1973). This technique is more sensitive than, and requires shorter
incubation times than the 02 light-dark bottle method. However, the 02 method
measures gross primary production, net primary production, and respiration, using
inexpensive laboratory reagents, while the 14C technique estimates only net primary
production and requires specialized training and equipment, as well as relatively
expensive radioisotopes.
Taxonomic Analysis
Subsamples drawn from water collected in water sampling bottles should be preserved
for later microscopic analysis to determine phytoplankton community composition. The
choice of fixation will depend on the dominant types of phytoplankton known to inhabit a
given area. (Buffered formaldehyde and Lugol's solution are two common fixatives.)
Preserved phytoplankton samples normally must be concentrated for quantitative
microscopic analysis. Analysis should include identification of dominant phytoplankton
taxa and counts of individual species. The description and comparison of phytoplankton
communities can be performed through the evaluation of species diversity, richness
(number of species), evenness, or numerous other parameters. Alterations to the
phytoplankton community should be analyzed in relation to other potential impacts on
other biological communities. These include but are not limited to:
• Food web impacts;
• Occurrence of toxic or nuisance phytoplankton; and
• Potential secondary impacts on zooplankton or fish communities (e.g., shift
from "more desirable" diatom-dominated communities to "less desirable"
flagellate- or cyanobacteria-dominated communities).
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Zooplankton
The taxonomic analysis should include the identification of dominant zooplankton taxa
and counts of individual species whenever possible. Particular attention should be given
to mesoplankton larvae of commercially, recreationally, or ecologically important
species. Characteristics of the zooplankton community should include but not be limited
to:
• Species composition (richness and evenness),
• Abundance,
• Dominance, and
• Diversity.
Alterations in the zooplankton community should also be analyzed in relation to potential
impacts on other biological communities. Such analyses should include, but not be
limited to, the structure and function of both larval and adult zooplankton communities
and consideration of food web impacts.
Biological Indices
The numbers of individuals and the numbers of species have been used as indicators of
anthropogenic disturbance, as well as other environmental stresses (USEPA, 1985b).
Furthermore, these simple biological indices are less ambiguous and can often be as
informative as diversity indices (USEPA, 1985b; Green, 1979; Hurlbert, 1971).
Measures of biomass have inherent problems in the collection of the data, e.g., loss or
gain of weight due to preservative medium, drying times, and evaporative weight loss.
More complicated indices such as species diversity richness, dominance, and evenness
have found varying degrees of acceptance. Although diversity may not be able to
pinpoint a single problematic discharge in an area that contains many, it can be an
indicator of degraded water quality. Species diversity of phytoplankton (Patten, 1962)
and zooplankton (Copeland, 1966; Copeland and Wohlschlag, 1968; Odum et al., 1963)
communities is noted to increase with increasing distance from a pollution source.
Diversity indices, which are measures of the distribution of individuals among species,
have the following limitations (Green, 1984a):
• They often lack biological meaning.
• They are not robust empirical indicators of any important correlates of
environmental health.
• They do not incorporate information on form and function of resident species.
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• They are susceptible to biases associated with well-described taxa.
Because of this, these methods should be used in conjunction with other assessment
methods to help verify results.
The dominance index is a measure of the degree to which one or a few species
dominate the community. The dominance index, herein defined as the minimum
number of species required to account for 75 percent of the total number of individuals,
has been useful in describing community structure (Swartz et al., 1985). Its advantages
are that it is easily calculated, it does not assume an underlying distribution of
individuals among species, and it is statistically testable.
Indicator Species
The evaluation of the abundance of individual indicator species is generally informative
and may reduce the cost of the analysis. The absence of pollution-sensitive species
and the enhancement of opportunistic and pollution-tolerant species may assist in
defining the spatial and temporal extent and magnitude of impacts. However, indicator
variables must possess the following characteristics (Green, 1984a):
• They must provide a sufficiently precise and accurate appraisal of:
- species of concern,
- anthropogenic disturbances to communities, and
- presence/absence or magnitude of anthropogenic perturbance to the
ecosystem.
• They must be a cost-effective and a statistically reliable alternative to
monitoring all critical measures of habitat perturbance.
• They must be appropriate for the spatial and temporal scale demanded by
the study objectives.
Further studies of the response patterns of zooplankton species subjected to
anthropogenic perturbations caused by discharges are required in order to select
appropriate indicators of environmental impact.
4.8.4
QA/QC Considerations
Variability in measurements caused by field heterogeneity is quantitatively determined
by the analysis of replicate field samples. Replicate sampling should be conducted at all
field stations where measurements are to be used in comparisons. Analysis of replicate
sample data is necessary for assessing the reliability of such comparisons.
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Laboratory performance and calibration should be verified at the beginning and
periodically during the time analyses are performed. Commercially available chlorophyll
is available and recommended for use in calibration. Chlorophyll quality control samples
are available from EPA's Environmental Monitoring and Support Laboratory in
Cincinnati, Ohio. Blind, split, or other control samples can be used to evaluate
performance. The Interim Guidance on Quality Assurance/Quality Control (QA/QC) for
the Estuarine Field and Laboratory Methods (USEPA, 1985b) provides a standard
operating procedure for chlorophyll measurements.
4.8.5 Statistical Design Considerations
Temporal stratification of the data should not be attempted until sufficient knowledge of
long-term natural cycles is attained. Initially, simple regression analyses may be
conducted on seasonally stratified data to identify monotronic temporal trends. Further
examinations of whether conditions are improving or degrading over time may be
conducted using statistical time series analyses (e.g., temporal autocorrelation, spectral
analyses, etc.).
Information derived from water quality monitoring will be important in interpreting the
results of plankton sampling. For example, phosphate concentrations may indicate the
cause of a phytoplankton "bloom," while measures of dissolved oxygen levels may
describe the effects of the "bloom" on other living estuarine resources. Therefore, the
selection of water quality and plankton sampling strategies must not be done
independently. These programs should be integrated to the fullest extent possible to
allow correlation of observed responses to changes in water quality parameters. Also,
alterations to the plankton community should be analyzed in relation to other impacts on
biological resources such as food web impacts on fish communities.
4.8.6
Use of Data
The analyzed, collected data will be used initially to identify the prominent species and
relate the temporal and spatial distribution patterns to water quality parameters.
Statistical techniques will be used to make these comparisons. As the discharge
guidelines are implemented, a plankton monitoring program can be used to track the
effectiveness of maintaining nutrient levels. As plankton levels fluctuate, their effects on
oxygen levels and fish and shellfish communities can be assessed. Plankton monitoring
strategies should be able to delineate between natural variability in plankton stocks and
variations caused by anthropogenic changes in nutrient concentrations.
Characterization of phytoplankton species abundance, distribution, and primary
productivity provides indications of water quality conditions. Monitoring changes in
phytoplankton community composition and densities is critical for the interpretation and
evaluation of long-term trends in water and habitat quality. Further understanding of the
causes of excessive water column and sediment oxygen demand requires tracking of
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photosynthetic activity and metabolic rates over time (Chesapeake Executive Council,
1988b). Zooplankton abundance and distribution are affected by both changes in
phytoplankton and changes in predator populations. Therefore, population
characteristics of this group can indicate symptoms of water quality problems, fishing
pressure, and other habitat problems for predator species (Chesapeake Executive
Council, 1988b).
4.8.7 Summary and Recommendations
Rationale
• Track phytoplankton and herbivore populations if changes in trophic structure
are suspected.
• Purpose of monitoring is to assess the effectiveness of the management
programs in mitigating the potential impacts caused by excessive nutrient
inputs and resulting changes in the planktonic community biomass and
structure/function.
Monitoring Design Considerations
• Collect both phytoplankton and zooplankton samples at specific depths
throughout the water column.
• Phytoplankton samples should be taken with water bottles and pumps used
for chlorophyll concentrations.
• Zooplankton samples should be taken with bottles or nets of varying sizes.
• Monitoring programs should consider natural and temporal fluctuations in
plankton biomass and species composition.
• Other components of the overall monitoring program, including water quality
monitoring and fish and shellfish communities, should be analyzed relative to
plankton community data to establish relationships and trends.
• Selected biological indices should retain biological meaning, be robust
indicators of marine "health," and incorporate species form and function.
• Indicator species should possess the following characteristics:
- sensitive to planktonic perturbances of concern;
- cost-effective and statistically reliable alternative to measuring all species
in a monitoring program;
- statistically reliable indicative measures of habitat perturbance; and
- appropriate for the spatial and temporal scale determined by the study
objectives.
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• Selection of reference sites is key to the evaluation of environmental impact
due to anthropogenic perturbances; several reference sites may be requried
to provide proper control for sampling sites.
Analytical Methods
• It is recommended that consistent types of sampling gear and location and timing
of sample collection be implemented to allow for comparisons between studies.
• Phytoplankton biomass can be determined using fluorometric or
spectrophotometric methods using a variety of filtration and extraction techniques.
• Phytoplankton productivity should be measured using the 14C light-dark bottle
technique.
• Taxonomic analysis should include identification and counts of dominant species.
QA/QC Analysis
• Replicate samples should be taken at all field stations where applicable.
• Laboratory performance evaluations and calibrations must be done on a
regularly scheduled basis.
• Standard chlorophyll samples, obtained from EPA, should be used for
calibrations.
Statistical Design Considerations
• Power analyses may be applied to determine the appropriate number of sample
replicates required to detect a specified difference.
• Temporal integration of the data should not be attempted until sufficient
knowledge of long-term natural cycles is attained.
Data Use
Identify dominant species.
Detect short- and long-term spatial and temporal trends in overall biomass and
productivity.
Detect short- and long-term spatial and temporal trends in species abundance,
distribution, and composition.
Examine relationship between water quality condition and trends in plankton
community characteristics.
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Examine relationship between plankton community characteristics and impacts
on other living resources (e.g., fish and shellfish communities).
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4.9
HABITAT IDENTIFICATION METHODS
Sensitive marine habitats serve a vital purpose as spawning grounds, nursery grounds,
forage areas, energy sources, shelter, and migratory routes and destinations for a
multitude of marine organisms. Pollution-related damage can cause long-lasting harm
to these habitats to the extent of altering ecosystem diversity and function.
For the purpose of this discussion, a sensitive habitat can be defined as a physical area
used, at least for a significant portion of the year, by a disproportionate abundance of
individuals and/or species, or as an area essential to the functioning of the ecosystem.
Habitats need not be permanent entities but may expand and contract on a seasonal
basis. Examples of transitory habitats include sea grass beds and ice floes. Marine
habitats can be organized in a hierarchical structure to be considered under the
discharge guidelines and can be grouped in the following manner:
Subtidal
Rock bottom
Bedrock/rubble
Coral
Oyster/clam shell reef
Unconsolidated bottom
Cobble/gravel
Sand
Mud-rich organic
Submerged aquatic vegetation (SAV)
Sea grass (Zostera, Thalassia)
Algal and other submerged plant communities
Water column
Intertidal
Emergent vegetation (along exposed coasts)
Marshes
Mangroves
4.9.1
Ice Floes
Rationale
The ocean discharge guidelines that relate to the presence of and impact on marine
habitats are the following;
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• Importance of the receiving water area to the surrounding biological
community, e.g., spawning sites, nursery/forage areas, migratory pathways
and areas necessary for critical life stages/functions of an organism;
• The existence of special aquatic sites, including (but not limited to) marine
sanctuaries/refuges, parks, monuments, national seashores, wilderness
areas, coral reefs/seagrass beds;
• Existing or potential recreational and commercial fishing; and
• Any applicable requirements of an approved CZMP.
The identification and delineation of these habitats should be made in both temporal and
spatial dimensions to account for variations in use by the biological community
throughout the year. Both the aerial extent and the functional value to biological
resources must be determined in considering the identification and importance of a
sensitive marine habitat.
Even though an effluent discharge may not directly alter the biological population in a
certain area, if the habitat essential for the survival of those individuals is destroyed, the
ecological results could be the same as if the population had been exposed directly.
The loss of critical habitat could cause the supported population to be either removed or
displaced to less suitable habitat.
Monitoring of discharge effects can help to ensure that long-term sublethal doses do not
accumulate and become toxic. It can also provide assurance that effects such as
increased turbidity or sedimentation do not reduce biological resource productivity below
sustainable levels.
During the permit review process, all important and sensitive habitats should be
identified and delineated to serve as a baseline for future monitoring activities. Habitat
quality in terms of functional values for fish, bird, and marine mammal populations must
also be taken into account to gain a complete valuation of the resources that could be
affected by discharges. Some habitat monitoring activities overlap with other biological
monitoring activities described in other sections. For habitats, the main concern is to
map the physical boundaries and to correlate biological populations to changes in
habitat distributions and functional values.
4.9.2 Monitoring Design Considerations
Habitat monitoring is a two-part process. First, the aerial coverage and boundaries must
be determined; second, the importance or the functional value to biological resources
must be determined. The first determination can be made by using charts, remote
sensing, or submersible remotely operated vehicles. The second analysis is carried out
by using quantification techniques to establish the value of that habitat for particular
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. Section 403 Procedural and Monitoring Quittance
species. The combining of these two techniques allows the identification of those
habitats critical to the functioning of the ecosystem. The determination of the functional
value of a habitat is much more subjective than delineation and must be done with care.
The results of these procedures should be combined with data on trends in species
distribution and water quality to provide a more complete assessment of the causes and
potential effects of habitat degradation. Habitat monitoring will also indicate which
environmental parameters and areas are critical to the ecosystem so that monitoring
activities can be tailored to assess these needs.
To conduct a habitat trend analysis, the sampling design must account for natural
variations caused by extraordinary meteorological events and other phenomena. The
recovery rate of habitat functions and its effects on the functional analysis described
above must also be considered. The primary concern of the habitat assessment is to
accumulate an adequate set of baseline data from which trends can be established.
During this initial phase, seasonal variations must be taken into account during sampling
phases. Sampling and delineation should be carried out at times of maximum biomass
or physical extent. For example, SAV beds or ice floes go through seasonal declines (or
disappearances) and thus a finite period of time exists to conduct sampling. Diurnal
factors such as tides and weather conditions also affect biological populations, and
sampling must be conducted taking these factors into account.
4.9.3 Analytical Methods Considerations
Habitats are generally classified in a hierarchical scheme as mentioned above.
Classification methods that can be used are referenced in Appendix A. The first course
of action in identifying, delineating, and evaluating critical marine habitats entails
mapping out the boundaries of the habitat, while the second involves a determination of
its value as an ecological resource. Table 4-23 lists the methods that can be used to
identify and evaluate sensitive habitats.
Table 4-23. List of Analytical Methods
Areal Trends
Functional Trends
Aerial Photography
Maps
Satellite Imagery
Remotely Operated Vehicles
Habitat Evaluation Procedure (HEP)
Habitat Suitability Modeling
Minimum Habitat Matrix
Wetland Evaluation Technique (WET)
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Methods
Habitat Delineation
Aerial Photography
Various government agencies (NOAA and USGS among others) have used aircraft to
produce aerial photographs of wide areas for a number of years. The use of
photographs in delineating marine habitats is of limited value since many features below
the low tide line cannot be viewed. Aerial photographs can be produced in a large range
of scales and spatial resolutions depending on the area in question. Fine detail can be
resolved for relatively small tracts, while large areas can be viewed at lower resolutions
to define gross features. In some regions there is an extensive record of historical
photographs that can be used to spot trends and to establish a baseline. New
overflights can be expensive and require ideal weather conditions, but they can be of
value for intertidal mapping and for mapping some submerged aquatic vegetation and
other shallow-water habitats.
Maps
As a preliminary method, maps and charts can be used to obtain a quick assessment of
the presence and delineation of sensitive habitats in the discharge area. Maps are the
least expensive spatial data source and are also the most simple. Nautical charts,
compiled by the Coast and Geodetic Survey of NOAA, contain bottom characteristic
information that could indicate the location of areas of biological importance. This
information is generally accurate for shallow-water areas but may be fragmentary or
nonexistent for areas farther offshore. The coordinated use of U.S. Geological Survey
topographic quadrangle maps, U.S. Fish and Wildlife Service National Wetlands
Inventory maps, and NOAA nautical charts can serve as an information base.
The principal advantages of using maps for delineating habitats is their ease of use and
their low cost. Disadvantages include the inaccuracies in the plotting of habitat
boundaries and the fact that the maps in use may be outdated and thus do not reflect
current conditions. In locations that have a history of human use, there may be an
extensive list of earlier chart or map editions.' These earlier editions can be an excellent
resource for spotting changes in habitat structure.
Satellite Imagery
Satellite imagery is a potential source of data when very large study areas are evaluated
and fine spatial detail is not required. This technique is useful to a maximum depth of
approximately 30 meters, and habitats can be resolved only to a minimum of roughly 30
to 50 square meters. For large study areas, the cost of satellite images can be high,
especially when a long-term temporal series may be required for monitoring purposes.
If the area under investigation contains large homogenous habitats, the resolution size
may not be a constraining factor. A major benefit of these images is that the spectral
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Section 403 Procedural and Monitoring Guidance
properties and brightness values are recorded digitally. This technology lends itself to
advanced spatial analysis methods that can be used to quantify productivity values.
These determinations are not possible with most aerial photographs.
Remotely Operated Vehicles
Remotely operated vehicles (ROVs) are small, unmanned, remotely operated
submersibles that are equipped with lights and cameras used to view portions of the
ocean bottom. These submersibles receive power and commands from a support
vessel and are highly versatile for identifying possible habitats. EPA Region 4 has used
ROVs to investigate and record the ocean bottom immediately'surrounding the
discharge outfalls of oil and gas drilling wells. This methodology produces a precise,
nondestructive record of the nature and extent of any physical habitat. The main liability
of this system is that unless vast amounts of time are spent, only small areas can be
investigated. These systems are also rather expensive and require well-trained
technicians to operate them. In addition, a ROV is limited by the length of its umbilical
and by the visibility of the water column in the search area. The visibility problem may
be compensated for by the addition of such technology as side scan sonar or sonar
attached to the ROV. Because of these constraints, the use of this system would be
confined to areas of most concern near the zone of initial dilution.
Functional Analysis
Functional analysis attempts to assess the use of the habitat by biological resources and
to compare the value of one habitat to that of another. Another aspect of functional
analysis is to calculate the change in a habitat's value over time. These processes
entail site-specific studies that can involve a large amount of time and money. The data
required by this technique can be collected in connection with the techniques described
in the biological methods section of this report. The analysis of these data will center on
spatial and temporal variations at a particular habitat. Various methods for conducting
functional analysis are referenced in Appendix A. Some of the most widely used
procedures are described below.
Habitat Evaluation Procedures
Habitat evaluation procedures (HEPs) are procedures developed by the U. S. Fish and
Wildlife Service to document the quality and quantity of available habitat for selected
wildlife species. The data generated are used to compare two habitats at the same
point in time and a single habitat over a length of time. By combining these two sets of
information, the impact of proposed discharges can be quantified (Lonard and Clairain,
1986).
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Methods
In the evaluation procedure, the habitat quality of selected species is documented based
on an evaluation of the ability of key habitat components to supply the life requisites of
the selected species. The evaluation involves using the same key habitat components
to compare existing habitat conditions and the optimum conditions for the species of
interest (USFWS, 1980).
The limitations in using this approach involve the use of habitat quality as an evaluation
parameter, which limits the methodology to those situations in which measurable and
predictable habitat changes are an important variable. This approach also forces a
long-term averaging type of analysis. There is no assurance that populations will exist
at the levels predicted by the analysis since all the environmental or behavior variables
that affect population levels may not be included. In addition, socioeconomic or political
constraints not taken into account may prevent the actual populations from reaching
their predicted levels (USFWS, 1980).
Habitat Suitability Modeling
Habitat suitability modeling consolidates habitat use information into a framework
appropriate for field application and is scaled to produce an index between 0.0 (unsuitable
habitat) and 1.0 (optimum habitat). Habitat suitability models are designed for use with the
habitat evaluation procedures described above (Toole et al., 1987). Multiple regression
equations are used to describe relationships between environmental characteristics and
productivity. Once the model has been validated, sensitivity analysis can be carried out to
project the results of varied discharge rates on the health of the site.
For modeling to make reliable predications, the input data must accurately reflect the
range of natural conditions observed. This may involve expensive and time-consuming
field studies in which the error limits and the QA/QC parameters must be well defined.
The reliability and the extent of the field data can be a major limitation in these studies.
Minimum Habitat Matrix
Minimum habitat guidelines for various species are a technique developed with the
ultimate goal of reestablishing balanced ecosystems in environmentally stressed areas.
This method is designed to provide information on the minimum habitat quality needed
by a target species and identifies those factors (both environmental and ecological)
required for the species. This information is formatted into a habitat requirement matrix
that defines the habitat parameters needed for successful reproduction and survival of
the indicated species. Such matrices can indicate the vital environmental parameters
that should be monitored, thus facilitating the development of monitoring programs. This
process is used to estimate the feasibility, benefits, and potential costs of maintaining
and protecting an estuarine environment suitable for the successful reproduction and
survival of the indicated species (Chesapeake Executive Council, 1988a).
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Section 403 Procedural and Monitoring Guidance
One potential problem with this method is that the primary indicated species and those
organisms on which the target species depend for food are both tracked with the
intention of maintaining habitat quality for both. This method may not completely
recognize the complex species interdependence within marine environments.
Wetland Evaluation Technique
The wetland evaluation technique (WET) was initially developed by the Federal Highway
Administration and revised by the U.S. Army Corps of Engineers to assess the function
and value of various components of habitats and their suitability for specific fish and
invertebrate species. WET evaluates functions and values in terms of social
significance, effectiveness, and opportunity (Adamus et al., 1987). Social significance
assesses the value of a site to society due to its special designation, potential economic
value, and strategic location. Effectiveness assesses the capability of a wetland to
perform a function as a result of its physical, chemical, or biological characteristics.
Opportunity assesses the opportunity for a habitat to perform a function to its level of
capability.
WET assesses functions and values by characterizing a habitat in terms of its physical,
chemical, and biological processes and attributes (Adamus et al., 1987). This
characterization is accomplished by identifying threshold values for predictors.
Predictors are simple, or integrated, variables that directly or indirectly measure the
physical, chemical, and biological processes or attributes of a habitat and its
surroundings.
4.9.4 QA/QC Considerations
Since sensitive habitat identification is relatively subjective, close controls must be
exercised during sampling and delineation. In the delineation of habitats, two individuals
should independently define the boundaries of a single habitat area. Any disagreement
between the two should be rectified before a final determination is reached. Sampling
programs demand a high level of expertise on the part of the individuals conducting
them. All personnel must have an advanced understanding of such topics as taxonomy,
sampling techniques, and statistics. The sampling strategy used will depend at least in
part on the level of expertise or training available. The selection of a functional analysis
method should be made after a review of the strengths and limitations of each method
has been conducted.
4.9.5 Statistical Design Considerations
In designing a habitat monitoring program, the researcher must first determine the
resources available. This will determine the extent of the monitoring program and the
level of detail it can provide. Aerial assessment can be done with limited resources,
while functional analysis requires greater commitments of both finances and time.
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Methods
Functional assessments require that the researcher decide which functions are of
greatest interest in the area of study or which will provide the most information
concerning the habitat attributes perceived to be of greatest value to the area.
Most habitat evaluation techniques involve some form of ranking, either quantitative or
qualitative. The purpose of the monitoring program is to periodically compare baseline
conditions to conditions after the permit has been issued. In terms of habitat monitoring,
this means comparing the baseline acreage or functional rating of the habitat to
conditions some time after discharge has commenced. Therefore, the ability to replicate
results is imperative.
4.9.6 Use of Data
The object of the data set is to obtain a clear and precise spatial and temporal picture of
the physical extent of a sensitive habitat. An evaluation of the functional value of that
habitat is also made so that the effects of various impacts can be judged. Sensitive
habitats should be clearly delineated on nautical charts or topographic maps along with
a complete description of their composition. The description should include a
characterization of the extent of natural seasonal variations in the boundaries. A
complete description of the functional value of the habitat to the biological community,
as well as an estimate of potential for loss of the resource from the effects of effluent
discharge, should also be made.
These data will act as a baseline for future monitoring programs that will track the
changes to the system. The baseline study and the monitoring program should be able
to discriminate between natural variations and discharge-induced changes.
4.9.7 Summary and Recommendations
Rationale
A sensitive habitat can be defined as a physical area utilized, at least for a
significant portion of the year, by a disproportionate abundance of individuals
and/or species, or as an area essential to the functioning of the ecosystem.
Sensitive marine habitats should be identified and their boundaries delineated
as part of an ongoing effort to assess the potential for impacts from effluent
discharges on ecological communities.
Monitoring programs that assess both the physical boundaries of the habitats
and the support functions that they serve for biological communities should
be conducted when sensitive habitats are present.
Damage to habitats from discharge effluents can cause irreversible harm and
create severe changes in ecosystem function and diversity.
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Section 403 Procedural and Monitoring Guidance
• Monitoring programs can provide assurance that effects such as increased
turbidity or sedimentation do not reduce biological resource productivity
below sustainable levels.
Monitoring Design Considerations
• Habitat monitoring consists of two parts: (1) aerial coverage and boundary
delineation and (2) functional analysis, which determines the value of the
habitat to the system.
• Habitat delineation can be fairly straightforward, but functional analysis is
rather subjective and must be done with care.
• Sampling strategies must account for variations caused by natural processes
and must be done at times of maximum extent or biomass.
• A functional analysis technique should be selected after a review of the
strengths and limitations of each method for each particular site.
Analytical Methods Considerations
• Sensitive habitats can be identified using maps, charts, and photographic
techniques. ROVs can be used to investigate near-field habitats in the zone
of initial dilution.
• Functional analysis determines the importance of a habitat to the biological
community.
• Functional analysis techniques are based on theoretical relationships of the
ecosystem; consequently, both the theory and the data used should be
vigorously evaluated.
• For habitat identification and delineations, the following techniques can be
used:
- aerial photography,
- maps/charts,
- satellite imagery, and
- remotely operated vehicles.
• Functional analysis can be carried out using various methods. Those
described are the following:
- habitat evaluation procedures (HEP),
- wetland evaluation technique (WET),
- habitat suitability modeling, and
- minimum habitat matrix.
773
image:
Methods
QA/QC Considerations
• Personnel identifying habitat should have advanced training, and two people
should independently define the boundaries.
Study Design Considerations
• Ability to replicate results is important.
• The extent of the monitoring program can be based on funding level.
• Aerial assessments are inexpensive, while functional analyses are more
expensive and more labor-intensive.
Use of Data
• Data collected will consist of charts identifying the boundaries of sensitive
habitats accompanied by complete descriptions.
• Functional analysis results will indicate which habitats are critical and the
effects observed under various discharge scenarios.
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Section 403 Procedural and Monitoring Guidance
4.10 BIOACCUMULATION
Bioaccumulation is the overall process of biological uptake and retention of chemical
contaminants obtained from foods, water, sediments, or any combination of exposure
pathways. A number of expressions are used to describe the bioaccumulation potential
of contaminants. These terms are explained in Table 4-24. Bioaccumulation is a
consequence of an organism's physiological limitations to transform and excrete the
invading chemical substances. Potential consequences of bioaccumulation of chemical
contaminants in marine organisms include, but are not limited to:
• Significant mortality to susceptible organisms;
• Lethal or sublethal chronic toxic responses at later stages of the life cycle or
under conditions of stress for susceptible organisms; and
• Transference of increasingly greater concentrations of toxic pollutants to
organisms at higher trophic levels, including humans.
Bioaccumulation analyses provide the link between exposure and effects and thus can
generate important insights into ecological effects, human health risks, and routes and
extent of pollutant exposure.
4.10.1 Rationale
The Clean Water Act section 403 point source discharge program requires (among other
things) a determination of:
Bioavailability
Bioaccumulation
Bioconcentration
Biomagnification
Table 4-24. List of Terms
A measure of the physicochemical access that a toxicant has to the
biological processes of an organism. The less the bioavailability of a
toxicant, the less its toxic effect on an organism (USEPA, 1991c).
Uptake and retention of substances by an organism from its surrounding
medium and from food (USEPA, 1990a).
Uptake of substances by an organism from the surrounding medium through
gill membranes or other external body surfaces (USEPA, 1990a).
The process by which the concentration of a compound increases in different
organisms occupying successive trophic levels (USEPA, 1990a)
175
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Methods
• The quantities, composition, and potential for bioaccumulation or persistence
of the pollutants to be discharged;
• The potential transport of the pollutants by biological, physical, or chemical
processes;
• The composition and vulnerability of potentially exposed biological communities;
and
• Potential direct or indirect impacts on human health.
Each of these determinations is addressed to some degree through bioaccumulation
studies.
Monitoring of bioaccumulation data is essential in relating the presence of selected
chemical residues in marine waters and sediment to their transfer and accumulation in
marine organisms and potential transfer to humans. Toxics can occur in the water
column at or near analytical detection limits; however, over time, these contaminants
can accumulate in fish and shellfish tissue to measurable concentrations. In areas
where water and sediments show measurable contaminant levels, it is difficult at best to
predict rates of uptake and bioaccumulation, although direct monitoring can provide
spatial and temporal records of the concentration and bioaccumulation of toxics. The
assessment of bioaccumulation in the surrounding biological community should be a
component of a monitoring program required in a permit issued under the "no
irreparable harm" clause of section 403.
4.10.2 Monitoring Design Considerations
Typically, bioaccumulation studies are formulated according to the study objectives. If
the objective is to determine the effects of bioaccumulation on human health, then
commercial and/or recreational fish and shellfish are tested. If determining the effects
on the habitat or ecosystem is the objective, usually the benthic macroinvertebrates are
surveyed. Unfortunately, comparisons between studies are valid only if the study
designs and procedures are comparable. Standardization of monitoring design would
allow for comparison between various studies.
A common deficiency in many programs is the inability to collect sufficient tissue
biomass of appropriate species across sampling locations throughout the study. The
selection of appropriate species and tissues must account for natural fluctuations in
populations as well as changes due to anthropogenic perturbations. Indigenous species
initially present may not be available later, limiting temporal and spatial comparisons.
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Section 403 Procedural and Monitoring Quittance
Selection of Target Species
A key component of any bioaccumulation monitoring program is the selection of target
species. Concentrations of chemical residues in tissues of target species serve as
indicators of contamination throughout the biological system. The fundamental criterion
is the ability to use the selected species to make comparisons between sampling
locations and sampling periods.
Target Species Characteristics
It is recommended that the target species possess the following characteristics (USEPA,
1985a):
• High bioaccumulation potential for selected contaminants of concern;
• Weakness or absence of metabolic regulation of selected contaminants to
allow assessment of a "worst-case scenario";
• Abundant enough, temporally and spatially, to allow for adequate sampling;
• Large enough to provide adequate amounts of tissue for analysis;
• Sessile or sedentary in nature to ensure that bioaccumulation is
representative of the study area;
• Easily collected; and
• Part of an existing data base of exposures and sensitivity.
Suggested fish and macroinvertebrate target species for various regions of the United
States are listed in Tables 4-25 and 4-26 (USEPA, 1985a).
Secondary considerations based on economic importance and status as a bioassay
organism may be applied to further winnow the list of candidate target species to a
practical number of species to be analyzed.
Candidate target species should be objectively ranked based on ecological
characteristics that would enhance their potential for bioaccumulation and facilitate
sampling and analytical procedures (USEPA, 1985a). The advantages of using benthic
macroinvertebrates as target species are:
• They are sendentary and therefore represent bioaccumulation in the study
area.
• They represent a significant food source for higher trophic levels and
therefore indicate the existence of contaminants in the food web.
177
image:
Methods
B
Table 4-25. Highest-Ranking Candidate Fish for Use
as Bioaccumulation Monitoring Species
Secondary Selection Criteria
State
MASSACHUSETTS
RHODE ISLAND
NEW YORK
NEW JERSEY
VIRGINIA
lE.iir' ^rjaSssi
Locality
Swampscott
Lynn
South Essex
Boston
Fall River
New Bedford
Newport
Upper East River
Lower East River
Lower Hudson River
Cape May
Portsmouth
Virginia Beach
1
Species
Winter flounder
Yellowtail flounder
Ocean pout
Windowpane
Winter flounder
Yellowtail flounder
Ocean pout
Winter flounder
Yellowtail flounder
Windowpane
American eel
Ocean pout
Winter flounder
Yellowtail flounder
Ocean pout
Windowpane
Winter flounder
Windowpane
Winter flounder
Scup
Summer flounder
Winter flounder
Scup
Weakfish
Winter flounder
Windowpane
Weakfish
Spot
Scup
American eel
Hogchoker
Spot
Red hake
Windowpane
Summer flounder
Spot
Summer flounder
Atlantic croaker
Hogchoker
Spot
Red hake
Summer flounder
tea *Jy,'XS >£«""!'»''*••* W»x A
eSst&'r XX^^X' .X •.'•£<^«*^'. "y«"*<iWfiS
Economic
Importance
Yes
Yes
No
No
Yes
Yes
No
Yes
No
No
No
No
Yes
Yes
No
No
Yes
No
Yes
No
Yes
Yes
Yes
Yes
Yes
No
Yes
Yes
Yes
No
No
Yes
No
No
Yes
Yes
Yes
No
No
Yes
No
Yes
* , #*v"*< £$•/£>>>* < < SoftiSw&jjy*'-??'?* ,
Bioassay
Species
Yes
Yes
No
No
Yes
No
No
Yes
No
No
No
No
Yes
No
No
No
Yes
No
Yes
No
Yes
Yes
No
No
Yes
No
No
Yes
No
No
No
Yes
No
No
Yes
Yes
Yes
No
No
Yes
No
Yes
yfAsffiW t.-s'^W-v?*' -. '-., ,v ,-W,
Xi * v'r-v *'""&?£%&*.??£?*
178
image:
Section 403 Procedural and Monitoring Quittance
* s , ->-.-•/>*•: ^"fj, ',-w, ''3&'j;?3&!
State
CALIFORNIA
(Northern)
CALIFORNIA
(Southern)
WASHINGTON
(^'•^tff^ff, , %J1 ^J'jJ^"** -fr"" V'*'*' **X %'"'****'• > JVJ , V^-Jifc, ^^^^K.^VX'-SSvMX.Xvs^A^'v^v..
' > *% A^V •." ' ' ' ' ^"'-s i /' * rti ' ,* Vv * •>•*«''* ^»i?v ^ - •• ^-V"V " ' * V»t"A' o' 5 > * >£ b^pw- v;
Table 4-25. (continued)
Locality
San Francisco
Oakland
Monterey
Santa Cruz
Watsonville
Goleta
Santa Barbara
L.A. County
Orange County
Hyperion
Oceanside
Escondido
San Elijo
San Diego
Central Puget Sound
^fy^ t°'Vv- •",, "
Secondary Selection Criteria
Economic Bioassay
Species Importance Species
English sole
Pacific sanddab
Big skate
English sole
Starry flounder
Pacific staghorn
sculpin
English sole
Curlfin sole
English sole
English sole
Curlfin sole
Dover sole
Pacific sanddab
Longspine combfish
Spotted cusk-eel
English sole
Pacific sanddab
Dover sole
Curlfin sole
English sole
Dover sole
Pacific sanddab
English sole
Dover sole
Longspine combfish
Big skate
California skate
Dover sole
Blackbelly eelpout
Pacific sanddab
English sole
Dover sole
English sole
Pacific sanddab
Longspine combfish
English sole
Dover sole
Rock sole
Spotted ratfish
Rex sole
C-O sole
Yes
Yes
No
Yes
Yes
No
Yes
Yes
Yes
Yes
Yes
Yes
Yes
No
No
Yes
Yes
No
No
Yes
Yes
Yes
Yes
Yes
No
No
No
Yes
No
Yes
Yes
Yes
Yes
Yes
No
Yes
Yes
Yes
No
Yes
Yes
Yes
No
No
Yes
No
No
Yes
No
Yes
Yes
No
No
No
No
No
Yes
No
No
No
Yes
Yes
Yes
Yes
No
No
No
No
No
No
No
Yes
No
Yes
No
No
Yes
No
No
No
No
No
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image:
Methods
Table 4-26. Recommended Large Macroinvertebrate Species
for Bioaccumulation Monitoring
Region
Recommended Species3
Massachusetts to Virginia
Alaska to California
Florida, Virgin Islands, and Puerto Rico
Hawaii
American lobster (Homarus americanus)
Eastern rock crab (Cancer irroratus)
Hard clam (Mercenaria mercenaris)
Soft-shell clam (Mya arenaris)
Ocean quahog (Artica islandica)
Surf clam (Spisula solidissima)
Edible mussel (Mytilus edulis)
Spiny lobster (Panulirus interruptus)
Dungeness crab (Cancer magistei1)
Rock crab (Cancer antennarius)
Yellow crab (Cancer anthonyi)
Red crab (Cancer productus)
California mussel (Mytilus californianus)
Spiny lobster (Panulirus argus)
Spiny lobster (Panulirus penicillatus)
a Additional species that may occur at specific discharge sites and are considered acceptable
bioaccumulation monitoring species include the American oyster (Crassostrea virginica) and the Pacific
oyster (Crassostrea gigas).
A disadvantage of macroinvertebrates is the lack of adequate tissue biomass for
analyses. Adult fish are often used not only for their significant biomass, but also as a
measure of possible contaminants available to the human population. Fish, however,
are motile and may not be representative of the study area.
Ideally, studies should include samples of numerous species from many phyla. This
diversity will ensure the collection of bioaccumulation data for a wide spectrum of
contaminants and will indicate a species' potential to bioaccumulate those contaminants.
For example, oysters and other bivalves are ideal for monitoring bioaccumulation of
PAHs because of their limited ability to metabolically transform these contaminants. The
monitoring of a food chain could establish the potential for transport from one trophic
level to another. Species with extensive bioaccumulation data would be preferred. A
fundamental criterion is the ability to use the selected species to make comparisons
between sampling locations and sampling periods.
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Section 403 Procedural and Monitoring Guidance
Caged Indicator vs. Indigenous Species
The California Mussel Watch Program, the U.S. Mussel Watch Program, and the NOAA
Status and Trends Program have employed the use of both resident and caged
transplant mussels to monitor bioaccumulation of toxic chemicals over space and time
(Goldberg etal., 1978; Boehm, 1984; Ladd etal., 1984). Caged indicator target species
offer several advantages over indigenous species :
. The biology and ecology of these indicator species are usually well described.
• Descriptions of culture and/or maintenance of the organisms under laboratory
conditions are often available.
.• The method provides control of initial temporal and spatial variation of
individuals and/or biomass.
• The method allows the use of a specific age, size, and/or genetic stock.
• The method ensures on-site bioaccumulation.
By controlling for initial conditions, the rate of change of tissue contamination may be
calculated. In addition, the Long Island Sound National Estuary Program has shown
that performing chemical residue analyses on caged indicator species appears to be a
promising approach for identifying sources of pollution (USEPA, 1982c).
The use of a caged indicator species indicates only uptake due to exposure to the water
column, and unfortunately methods for caging sediment-ingesting organisms are not yet
available. The bioaccumulation of sediment-sorbed contaminants is thought to be
indicated by transfer through interstitial waters (Knezovich and Harrison, 1987), and the
results often do not relate to species found on site. Therefore, caged organisms may
not provide a full-spectrum profile of the contaminants that are bioavailable at any one
site.
The advantages of indigenous species are as follows:
• Results obtained will relate directly to those species which may be impacted
and will provide a direct measure of potential risks to human health.
• There is no cost of transplanting organisms.
• There is no possibility of loss of cage/buoy system.
• There is no possibility of introduction of a "nuisance" species.
Unfortunately, common indigenous species may not meet the criteria for use as
bioaccumulation target species, negating any advantage of using a native species. The
most common problem is collecting sufficient tissue biomass of appropriate species
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across sampling locations throughout the study. The composition of benthic community
species is subject to large natural fluctuations, as well as fluctuations in response to
pollution (Pearson and Rosenberg, 1978); indigenous species initially present may not
be available later, limiting temporal and spatial comparisons.
Selection of Tissues
The type and location of tissue analyzed will depend on the objectives of individual
monitoring programs. It is strongly recommended that all analyses be conducted with
the same species and tissue type in order that scientifically and statistically valid
comparisons may be made.
In fish, liver tissue is closely associated with regulation and storage of many toxic
substances (Fowler, 1982). Its high fatty tissue content tends to accumulate lipophilic
contaminants. Therefore, contaminant levels in the liver can be used to estimate the
range of contaminants being assimilated. For macroinvertebrates, hepatopancreas or
digestive gland tissue performs functions analogous to those of fish liver tissue.
Contaminants in edible muscle tissue represent those contaminants that are retained in
a form that allows transfer to humans. Sampling of muscle tissue is appropriate for
human exposure assessments and quantitative health risk determinations. Within a
fillet, contaminant concentrations may vary; therefore, it is recommended that a
consistent location within the muscle tissue be analyzed (USEPA, 1989a).
Whole-body analyses should be conducted when predators consume the whole body of
the target organism. If organisms are not cleansed of materials contained in the
digestive tract, contaminants in the gut contents will be included in the analyses. To
provide the most accurate estimate of the total amount of contaminants available to
most macroinvertebrate predators, this type of cleansing may not be required.
Total Organic Carbon and Acid Volatile Sulfides. and Tissue Lipid Normalization
Toxic sediment concentrations of hydrophobic contaminants have been found to be
related to the total organic carbon content (TOC) of the sediment (Karickhoff et al.,
1979). The bioavailability of metal contaminants has been found to be related to the
acid volatile sulfide (AVS) concentrations of the sediment (DiToro et al., in press). TOC
and AVS normalizations have been conducted to estimate the concentrations of
sediment contaminants that are bioavailable over different sampling locations.
Sediment TOC and AVS normalizations are recommended to account for the variability
in bioavailable sediment contaminant concentrations between locations.
Lipids appear to be a storage site of organochlorines, hydrocarbons, and other
hydrophobic contaminants in a variety of organisms (de Boer, 1988). In fact, just as
TOC and AVS are considered by many to be the most important parameters in defining
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organic and metal concentrations in sediments, lipid content is considered the salient
parameter in defining hydrophobic residue concentrations found in tissues (Phillips and
Segar, 1986). Likewise, tissue lipid normalizations have been suggested to account for
the variability in tissue contaminant concentrations between individuals.
Normalization of sediment contaminant concentrations to TOG and AVS and
normalization of tissue contaminant concentrations to the tissue lipid concentrations
have been made to permit comparisons of toxic chemical residues in tissues between
studies, locations, individuals, and tissue types. These sediment and tissue
normalizations assume the following::
• Contaminants partition predominantly to sediment TOC, sediment AVS, and
organismal lipids.
• Contaminants partition reversibly between the sediment particles and the
organism.
• Rapid steady-state kinetics of contaminants are maintained.
• Sediment is the only source for the bioaccumulation.
Lipid normalization has been performed in order that individuals with differing body fat
levels may be compared. Again, it is essential that all analyses be conducted with the
same tissue type from the same species to ensure that scientifically and statistically
valid comparisons may be made. TOC/lipid normalized accumulation factors (AF) have
also been used to predict tissue residue concentrations (Ferraro et al., 1990; Lake et al.,
1987).
Threefold differences in lipid concentration may result as a consequence of various lipid
analysis techniques. Development of protocol standardization or intercalibration
between lipid techniques is required before chemical residues in tissues can be
compared between studies, Microtechniques have been developed for analyses of
lipids from single individuals (Gardner et al., 1985). Finally, the decision to normalize for
TOC, AVS, and percent Ispid will depend on monitoring program objectives. For
example, if the monitoring objective is to identify "hot spots"—i.e., those areas where
high body burdens present a risk to human and ecological health—normalization may
not be appropriate. If the objective is to identify temporal trends in bioaccumulation
rates, however, normalization may be justified.
Selection of Sampling Period
The timing of sampling should be based on biological cycles that influence an
organism's susceptibility to bioaccumulation. For example, for crustaceans, just after
molting and before hardening of the integument occurs, restricting its permeability, there
is a significant increase in potential for bioaccumulation of toxic contaminants. The
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frequency of sampling should be related to the expected rate of change in tissue
concentrations of contaminants. A consistent sampling period is recommended in order
that spatial and temporal comparisons may be conducted.
For many aquatic vertebrate and invertebrate organisms, the reproductive cycle exerts a
major influence on tissue concentrations of many contaminants, especially lipophilic
compounds (Phillips, 1980). For many species, lipophilic contaminants are transferred
from the muscle and liver to the eggs as lipids are mobilized and transported to the egg
during oogenesis (Spies et al., 1988; Gardner et al., 1985). There is evidence that
lipophilic contaminants have deleterious effects on the developing egg and/or the larvae
(Spies et al., 1988; Hansen et al., 1985; Niimi, 1983). Transfer of contaminants from
adult to egg and its potential impacts on future fisheries recruitment and fisheries
production are under investigation.
It is recommended that target species be sampled when tissue contaminant
concentrations are expected to be at their highest levels in order to evaluate worst-case
scenarios. For those organisms where the body is consumed whole, contaminant levels
are usually at their highest at or just before spawning. For organisms where the muscle
tissue is consumed, contaminants in muscle tissue usually reach a peak well before
spawning.
4.10.3 Analytical Methods Considerations
Factors to be considered during the choice of an appropriate analytical method include
the parameters of interest, desired detection limits, sample size requirements or
restrictions, methods of preservation, technical and practical holding times, and matrix
interferences. Several USEPA documents (1986b, 1987c, 1990c) discuss the common
analytical problems encountered during monitoring analyses of tissue samples.
Chemical Residue Analyses
Several factors determine achievable detection limits for a specific contaminant,
regardless of analytical procedure. These factors include:
• Size of the tissue sample available for analysis. In general, the more tissue
available for analysis, the better the detection levels that can be achieved; a
minimum of 30 g (wet weight) is usually considered adequate (USEPA,
1987c);
• Presence of interfering substances;
• Range of pollutants to be analyzed—an optimal method may be developed
without regard to potential effects on other parameters;
• Level of confirmation—qualitative vs. quantitative analyses; and
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• Level of pollutant found in the field and in analytical blanks.
Selection of appropriate methods ought to be based on a trade-off between full-scan
analyses, which are economical but cannot provide optimal sensitivity for some
compounds, and alternative methods that are more sensitive for specific compounds but
can result in higher analytical costs. A list of analytical techniques is presented in Table
4-27.
The Computerized Risk and Bioaccumulation System (CRABS, Version 1.0) is an expert
system designed to predict tissue residues of 15 neutral organic pollutants in
sediment-dwelling organisms and the human cancer risk from consumption of
contaminated shellfish (USEPA, 1990c). Thermodynamic partitioning, first-order
kinetics, or toxicokinetic models are used to predict bioaccumulation from bedded
sediment. Steady-state tissue residues are predicted from any of the models, whereas
the two kinetic models are used to predict either non-steady-state uptake or elimination.
The lifetime human cancer risk is then predicted for the consumption of clams and other
nonmobile sediment-dwelling organisms containing the predicted or measured tissue
residue. The cancer risk is predicted for a single pollutant from a single-species diet.
Shellfish consumption rates can be estimated if site-specific rates are available.
Table 4-27. List of Analytical Techniques (USEPA, 1986b)
METALS/METALLOIDS
• Atomic Absorption Spectrophotometry (AA) USEPA 7000 series methods
flame
graphite furnace (GFAA)
cold vapor
gaseous hydride (HYDAA)
USEPA method 7470
USEPA methods 7060 and 7740
• Inductively Coupled Plasma Emission Spectrometry (ICP) USEPA method 6010
ORGANICS
• Gas Chromatography (GC)
with electron capture detection (GC/ECD)
with mass spectrometry (GC/MS)
USEPA method 8080
USEPA methods 8240 and 8270
SOURCE: USEPA, 1986b.
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This program is now available for distribution. All inquiries should be directed to the
Narragansett, Rhode Island, Environmental Research Laboratory (Hatfield Marine
Science Center, Newport, OR 97365, 503/867-4042).
Tissue residue data necessary for the evaluation of dredged materials and for ecological
and human risk assessments are most directly derived using bedded sediments (i.e.,
deposited rather than suspended sediments) in bioaccumulation tests (USEPA, 1989d).
Sediments are the largest receiving grounds for toxics known to bind with particles, such
as organics with high octanol-water partitioning coefficients (e.g., PCBs and DDT) and
many heavy metals. Previous techniques have varied to meet specific requirements for
the task at hand, and comparability of results is questionable. USEPA (1989d)
standardizes an approach for conducting sediment bioaccumulation tests using
sediment-ingesting organisms exposed to bedded marine sediments. Guidance is
presented for "routine" testing in the laboratory and has not been tailored for specific
regulations or geographic locations. Data can be generated by the 28-day test that is
applicable for quantitative ecological and human health risk assessments. It should be
noted that this is a "living" document subject to revisions warranted by experience.
Metals and Metalloids
Nitric acid/perchloric acid digestions should be considered for analysis of tissue
samples. Perchloric acid is especially useful for the dissolution of fat.
Trace element analyses by inductively coupled plasma emission spectrometry (ICP)
allow for several elements to be measured simultaneously. However, the detection
limits of ICP are generally not as sensitive as those achieved by graphite furnace atomic
absorption spectrophotometry (GFAA).
The combination of atomic absorption spectrophotometry techniques (AA) and ICP is
the recommended analytical method for detection of metals and metalloids since no
technique is best for all elements. Cold vapor AA analysis is the only recommended
technique for mercury.
GFAA is more sensitive than flame AA, but is more subject to matrix and spectral
influences. GFAA requires particular caution with regard to laboratory contamination. In
either case, the concentration of each element is determined by a separate analysis,
making the analysis of a large number of contaminant metals both labor-intensive and
relatively expensive compared to ICP.
Semivolatile Organic Compounds
Analysis of semivolatile organic compounds involves a solvent extraction of the sample,
cleanup of the characteristically complex extract, and gas chromatography (GC)
analysis and quantification. There are two gas chromatography/mass spectrometry
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(GC/MS) options for detecting extractable organic compounds: internal standard
technique and isotope dilution. Isotope dilution is recommended because reliable
recovery corrections can be made for each analyte with a labeled analog or a chemically
similar analog (Tetra Tech, 1986).
The identification of pesticides and PCB can be made by gas chromatography/electron
capture detection (GC/ECD) analysis. GC/ECD provides greater sensitivity relative to
GC/MS; however, GC/ECD does not provide positive compound identification.
Confirmation of pesticides and PCBs on an alternative GC/ECD column or preferably by
GC/MS, when sufficient concentrations occur, is recommended for reliable results (Tetra
Tech, 1985). All other organic compound groups are recommended for analysis by
GC/MS (Tetra Tech, 1985).
Volatile Organic Compounds
The purge-and-trap GC/MS technique is employed for detecting volatile organic
compounds in water. A successful variation for detection of volatile organic residues in
tissues involves a device that vaporizes volatile organic compounds from the tissue
sample under vacuum and then condenses the volatiles in a super-cooled trap (Hiatt,
1981). The trap is then transferred to a purge-and-trap device, where it is treated as a
water sample. The isotope dilution option is recommended because it provides reliable
recovery data for each analyte (Tetra Tech, 1986).
4.10.4 QA/QC Considerations
Analysis of blanks should be conducted to demonstrate freedom from contamination. At
least one method blank must be included with each batch of samples and must
constitute at least 5 percent of all samples analyzed.
Spike recovery analyses are required to assess method performance on the sample
matrix. This method serves as an indication of analytical accuracy, but not necessarily
of extraction efficiency.
Replicates must be analyzed to monitor the precision of laboratory analyses. A
minimum of 5 percent of the analyses should be laboratory replicates. Triplicates should
be performed with each sample batch over 40 samples.
Laboratory performance and calibration should be verified at the beginning and end of
each 12-hr shift during which analyses are performed.
Reports delineating the essential elements of the bioaccumulation component of the
program should be included with the quantitative QA/QC analyses. It is recommended
that these reports be recorded and stored in a data base for future reference.
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4.10.5 Statistical Design Considerations
Composite Sampling
Composite tissue sampling consists of mixing tissue samples from two or more
individual organisms, typically of a single species, collected at a particular location and
time period. The analysis of a composite sample provides an estimate of an average
tissue concentration for the individual organisms composing the composite sample.
Advantages of the composite sampling strategy are :
• It provides a cost-effective strategy when individual chemical analyses are
expensive.
• It provides a means to analyze bioaccumulation when the tissue mass of an
individual is insufficient for the analytical protocol.
• It results in a more efficient estimate of the mean at specified sampling
locations.
Because of the reduced sample variance, composite sampling results in a considerable
increase in statistical power. If the primary objective of a monitoring program is to
determine differences in contaminant tissue among sampling locations, composite
sampling is an appropriate strategy.
Composite sampling is not recommended if the objective of the monitoring program is to
determine compliance with specified tissue contaminant concentration limits since this
sampling method does not detect the true range of tissue contaminant concentrations in
the population. Special considerations related to composite sampling include :
• The range and the variance of the population of individual samples cannot be
directly estimated.
• If species are mixed, tissue composites are likely to be composed of different
proportions of species and numbers of individuals, confounding whether
patterns of tissue residue concentrations are due to differences in locations or
to interspecific differences in bioaccumulation.
Space-bulking consists of sampling of individual organisms from several locations and
combining tissue samples into one or more composite samples. Time-bulking involves
taking multiple samples over time from a single location and compositing these samples.
The use of space- and/or time-bulking strategies should be carefully considered since
significant information concerning spatial and temporal heterogeneity may be lost. The
adoption of composite strategies will depend on the objective of individual monitoring
programs.
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Statistical Power
Tetra Tech (1987) demonstrated that the statistical power increases with the increase of
the number of individuals in each replicate composite sample. However, a diminished
return of statistical power exists with the addition of successive individuals to each
composite. For composites of greater than 10 individuals, the increase of power is
negligible given typical levels of data variability.
For moderate levels of variability in chemical residue data, 6 to 10 individuals composing
each of 5 replicate composites should be adequate to detect a treatment difference
equal to 100 percent of the overall mean among treatments (Tetra Tech, 1987). For
analyses of individuals, a minimum of five organisms at each site is recommended.
To improve the power of a statistical test, while keeping the significance level constant,
the sample size should be increased. Because of constraints in cost and time, however,
this option may not be available. Power analyses have shown that for a fixed level of
sampling effort, a monitoring program's power is generally increased by collecting more
replicates at fewer locations. The number and distribution of sampling locations
required to evaluate the effect of a discharge will depend on the volume and transport of
the effluent.
4.10.6 Use of Data
Results of the bioaccumulation analyses can be used to :
• Establish spatial and temporal trends in the bioaccumulation of toxicants of
selected marine fish and macroinvertebrates;
• Identify existing arid potential problem areas for fish and macroinvertebrate
contamination; and
• Supply data that can be used to calculate the human health risk of
consuming marine fish and shellfish.
4.10.7 Summary
Rationale
• Monitoring the accumulation of chemical residues in tissues of marine
organisms will provide information essential in relating the presence of
selected contaminants in marine waters and sediment to their transfer and
accumulation in marine organisms.
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Monitoring Design Considerations
• Target species should possess the following characteristics :
- high bioaccumulation potential for selected contaminants of concern;
- weak or absent metabolic regulation of selected contaminants;
- abundant enough, temporally and spatially, to allow for adequate
sampling;
- large enough to provide adequate amounts of tissue for analysis;
- sessile or sedentary in nature to ensure bioaccumulation is representative
of the study area; and
- easily collected.
• Caged indicator species
- Method allows for control of initial temporal and spatial variation of
individuals and/or biomass.
- Method allows for the use of specific age, size, and/or genetic stocks.
• Indigenous species
- Results obtained will relate directly to those species that may be impacted.
• Target tissues
- Fish liver or macroinvertebrate hepatopancreas analyses may be used to
estimate the range of contaminants being assimilated.
- Muscle tissue analyses are appropriate for human exposure assessments
and quantitative health risk determinations.
- Whole-body analyses should to be conducted when predators consume
the whole body of the target organism.
• TOC/Lipid Normalization
- Allow comparisons of chemical residue concentrations in tissues between
locations, individuals, and tissue type.
- Limitations in use due to differences in lipid analysis techniques.
Time of Sampling
- Timing of sampling should be based on biological cycles that influence an
organism's susceptibility to bioaccumulate.
- It is recommended that target species be sampled when tissue
contamination concentrations are expected to be at their highest level.
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Analytical Methods Considerations
• Metals/Metalloids
- A combination of AA and IGP is recommended for the detection of metals
and metalloids.
- Cold vapor AA is the recommended protocol for mercury detection.
• Organics
- GC/MS in conjunction with isotope dilution is recommended for the
detection of semi volatile organic compounds.
- Super-cooled trap, in conjunction with a purge-and-trap device, is
recommended for the detection of volatile organics.
QA/QC Considerations
• Blank, spike recovery, and replicate analyses are recommended quality
control checks.
• Reports delineating the essential elements of the bioaccumulation
component of the program should be included with the quantitative QA/QC
analyses.
Statistical Design Considerations
• Compositing tissue sampling consists of mixing tissue samples from two or
more individual organisms collected at a particular location and time period.
• Space-bulking (combining composites from several locations) and
time-bulking (combining several composites over time from one location)
strategies should be used judiciously because information concerning spatial
and temporal heterogeneity may be lost.
• Six individuals composing each of five replicate composites should be
adequate to detect a treatment difference equal to 100 percent of the overall
mean among treatments.
Use of Data
• Establish spatial and temporal trends in the bioaccumulation of toxicants of
selected marine fish and macroinvertebrates.
• Identify existing arid potential problem areas for fish and macroinvertebrate
contamination.
• Supply data that can be used to calculate the human health risk of
consuming marine fish and shellfish.
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4.11 PATHOGENS
Monitoring for pathogenic microorganisms is currently conducted by State environmental
and human health agencies in shellfish harvesting areas and bathing beaches. National
Pollutant Discharge Elimination System (NPDES) pathogen monitoring programs may
be under way at selected locations both to assess the condition of water in the vicinity of
discharges and surrounding areas and to assess relative pathogen contributions from
these permitted effluent discharges.
The National Shellfish Sanitation Program (NSSP) has been established as a joint effort
involving the Food and Drug Administration (FDA), State agencies, and the shellfish
industry to set forth guidelines for the management of State shellfish programs. As part
of the NSSP, the Food and Drug Administration (FDA) provides technical assistance to
States for studying specific pollution problems, provides data to establish closure levels
for shellfish harvesting, conducts applied research on various contaminants to assist in
developing standards and criteria, and evaluates the effectiveness of State shellfish
sanitary control programs. In addition, since 1966, data have been compiled periodically
by FDA and the National Oceanic and Atmospheric Administration (NOAA) on the
classification by States of coastal and estuarine waters with regard to suitability for
shellfishing activities. In addition to classifying their waters as to their suitability as
shellfish harvesting areas, States also issue beach closures. These closures are
typically based on water quality criteria developed by the Federal government.
4.11.1 Rationale
Monitoring of pathogens in the marine environment can be used in the CWA section 403
point source discharge program to address the following ocean discharge guideline:
• Potential direct or indirect impacts on human health.
Human pathogens found in the marine environment include viruses, bacteria,
protozoans, helminth and parasites. In the United States, viruses and bacteria are the
most important pathogens, in terms of both the number of organisms released to the
environment and the severity of the diseases they cause (OTA, 1987).
Humans can be exposed to pathogens by direct contact with contaminated waters (e.g.,
swimming, surfing, diving) or indirectly through ingestion of contaminated food (e.g.,
molluscan shellfish). The preponderance of evidence indicates that the etiologic agents
for waterborne outbreaks of acute gastroenteritis (AGI) are the Norwalk-like viruses
(Kaplan et al. 1982). Hepatitis A has been linked to the consumption of raw or partially
cooked molluscan shellfish (Feingold, 1973; CDC, 1979; Ohara et al., 1983). Bacteria
responsible for typhoid and cholera are known to be water- and seafood-borne. These
pathogens can enter the marine environment through the discharge of raw sewage,
wastewater effluent from sewage treatment plants, failing septic tanks, and the dumping
of sewage sludge (OTA, 1987). Monitoring of human pathogens provides information
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essential in relating the presence of pathogens in marine waters and shellfish to
outbreaks of disease. Monitoring data may be used to identify potential sources of
pathogens. The assessment of pathogen contamination should be a component of a
monitoring program where pathogens may present a risk to human health and the
economic vitality of the marine environment.
4.11.2 Monitoring Design Considerations
Water Column Sampling
Bacteria are not uniformly distributed throughout the water column (Gameson, 1983);
bacterial abundances, several orders of magnitude greater than underlying waters, can
be found in a thin microlayer on the surface of the water (Hardy, 1982). If feasible, it is
recommended that samples of this microlayer and separate samples of the underlying
waters be collected. However, standardized methods for sampling this microlayer have
not been established. If it is not feasible to sample both the microlayer and underlying
waters, the "scoop" method should be used to ensure that the surface microlayer is
sampled (USEPA, 1978).
Water samples for bacterial analyses are frequently collected using sterilized plastic
bags (e.g., Whirl-Pak) or screw-cap, wide-mouthed bottles. Several depths may be
sampled during one cast, and/or replicate samples may be collected at a particular
depth by using Kemmerer or Niskin samplers (USEPA, 1978). Any device that collects
water samples in unsterilized tubes should not be used for collecting bacteriological
samples without first obtaining data that support its use. Pumps may be used to sample
large volumes of the water column (USEPA, 1978).
Sentinel Organisms
Analyses of sentinel shellfish tissue (e.g., mussels and oysters) offer several
advantages:
• They concentrate pathogens and may be useful viral analyses since viruses
are often present in low numbers.
• They provide a means of temporally integrating water quality conditions.
• They may be direct measures of human exposure to pathogens (i.e.,
consumption of contaminated food).
• They may be deployed and maintained in a number of diverse locales.
However, the disadvantages of sentinel organisms include:
• Transplanting organisms may be costly;
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There is a possibility of losing the cage/buoy system; and
Species may not be tolerant of all test site conditions (e.g., low salinity, high
turbidity).
It is essential that monitoring design elements be standardized to allow for comparisons
between marine studies. The selection of a standard sentinel species appropriate to the
locality is recommended; interspecies differences would not allow comparisons of body
burdens.
4.11.3 Analytical Methods Considerations
Many pathogens can be present in wastes, contaminated media, and infected
organisms. Human pathogens are of concern in marine waters because they are:
• Associated with major debilitating diseases (e.g., hepatitis, cholera);
• Infectious at low doses;
• Resistant to environmental stress; and
• Not readily enumerated due to low numbers.
Table 4-28 provides examples of pathogenic organisms known to cause adverse human
health effects. The alternative to directly identifying pathogens of concern has been the
enumeration of bacteria, which are indicators of human waste contamination or
indicators of human illness.
Indicators of Human Health Risks
Because of the inability to enumerate pathogens of concern, indicators of human
pathogen densities have been used to assess human health risks. Indicators useful in
predicting infectious disease rates should have the following characteristics:
• High abundances should be consistently found in human fecal wastes.
• No significant extra-human fecal sources should be present.
• The indicators must provide temporally and spatially reliable and accurate
appraisals of the pathogen of concern.
Currently, there is no consensus on which indicator organism, if any, is specific for
human feces.
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Table 4-28. Microorganisms Responsible for Causing Adverse Human Health
Effects3
Disease
Pathogenic
Organism
Seafood
Source
Hepatitis
Gastroenteritis
Hepatitis A virus
Non-A and non-B
hepatitis virus
Aeromonas hydrophilia
and Plesiomonas
shigelloides
Vibrio mimicus
Vibrio parahaemolyiicus
Vibrio vulnificus
Vibrio cholera, O group
Vibrio cholera,
Non-O group 1
Norwalk virus
Small round structured
virus
Campylobacterjejuni
Raw oysters
Steamed clams
Steamed and raw clams
Raw clams
Raw oysters
Oysters
Cockles
Raw molluscan shellfish
Raw molluscan shellfish
Shellfish
Raw oysters
Clams and snails
Raw oysters
Crab
Shrimp
Lobster
Raw oysters
Raw oysters
Boiled shrimp, boiled crab
Raw oysters
Raw oysters
Raw clams
Raw oysters
Raw clams
a Includes naturally occurring microorganisms as well as microorganisms associated with pollution.
SOURCE: NOAA, 1988
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Coliform Bacteria
For decades, the concentration of coliform bacteria, either total or fecal coliform, has
been considered a reliable indicator of the presence and density of pathogens. The use
of total coliform bacterial criteria to protect human health from disease associated with
contaminated water is widespread. Furthermore, total coliform concentrations have the
advantage of providing a basis for comparison with historical data (USEPA, 1988f).
However, these indicators and their corresponding water quality standards have not
been related to incidences of disease through epidemiological studies. Controversy
exists on the efficacy of coliform bacteria to predict the presence and inactivation of
other types of pathogens (Pederson, 1980). In fact, recent studies indicate that fecal
coliform bacteria may not be a reliable indicator for predicting the risks associated with
direct exposures to pathogens in the marine environment (Cabelli et al., 1979, 1982).
Fecal conforms are not pathogenic and are less resistant to environmental stress
compared to many pathogens (Borrego et al., 1983). Furthermore, fecal coliform
bacteria are not specific to mammalian fecal pollution. The lack of this specificity
prompted the development of methods for enumerating fecal coliform bacteria specific to
mammalian fecal pollution (e.g., Escherichia coli). However, with respect to recreational
water quality criteria, EPA has not recommended the use of E. coli tor marine waters.
Enterococci
Enterococci are streptococcus bacteria indigenous to the intestines of warm-blooded
animals. Cabelli et al. (1983) found that the densities of enterococci were highly
correlated with the incidence of gastrointestinal (Gl) symptoms among swimmers;
reported swimming-associated Gl symptoms were poorly correlated with fecal coliform
densities. Enterococci also have the following advantageous characteristics:
They are tolerant to high salinity and are of particular value in the analysis of
marine waters.
Taxonomic identifications are relatively simple and can reveal the origin of
mammalian pollution (e.g., humans, livestock).
Genetic fingerprinting techniques that can link these bacteria in the
environment to specific sources of contamination have been developed.
State-of-the-art genetic fingerprinting techniques are very costly and require further
study in order to assess their reliability. EPA has adopted enterococci as an indicator of
microbiological water quality for recreational marine waters.
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Viruses
Viruses are being recognized as major etiologic agents for many outbreaks of human
illness as a result of swimming in contaminated water and consuming contaminated
seafood. Viruses are excreted only by infected individuals; they are not normal flora in
the intestinal tract. Traditional indicators of microbial contaminants appear to be
inadequate for predicting human health risks associated with both consuming molluscan
shellfish and swimming in waters that contain sewage-associated viruses. However,
examination of water for enteric viruses is not recommended at this time, except in
special circumstances, because of methodology limitations. Even state-of-the-art
methods for concentrating viruses from water are still being researched and continue to
be modified and improved. None of the available virus detection methods have been
tested adequately with representatives from all of the virus groups of public health
importance. In addition, some of these methods require expensive equipment and
materials for sample processing and all virus assay and identification procedures require
expensive cell culture and related virology laboratory facilities.
Laboratory Techniques
It should also be noted that no single procedure is adequate to isolate all
microorganisms from water, and the presence of one microorganism does not signify the
presence or absence of any other (Table 4-29). A more detailed description of analytical
methods for a number of pathogens is presented in Standard Methods for the
Examination of Water and Wastewater (APHA, 1989).
Fecal Bacteria
Two standard methods are presented here for the detection of fecal bacteria: the
membrane filter procedure and the multiple-tube fermentation procedure.
The membrane filter (MF) technique involves sample filtration followed by direct plating
for detection and enumeration of coliform bacterial densities. The MF technique can be
used to test relatively large volumes of samples and yields numerical results more
rapidly than does the multiple-tube procedure. The statistical reliability of the MF
technique is greater than that of the Most Probable Number (MPN) procedures (APHA,
1989). However, the MF technique has limitations, particularly in testing waters with
high turbidity and noncoliforrn (background) bacteria. The MF technique can be used to
measure bacterial densities of Escherichia coll (E. coif) and enterococci in ambient
waters (USEPA, 1985e) since EPA has approved this technique for use in seawater.
The multiple-tube fermentation technique involves a series of fermentation tubes
containing growth media that are inoculated with the appropriate decimal dilutions of
water (multiples and submultiples of 10 ml_), based on the probable coliform density.
Following inoculation at 35 ± 0.5°C for 24 ± 2 hr, the tubes are examined for gas or
acidic growth (distinctive yellow color). If no gas or acidic growth has formed, the
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Table 4-29. Laboratory Procedures for Bacterial Indicators
Laboratory Procedures
Test Organisms
Fecal coliform
bacteria
Water
MPN tubes using A-l
broth (APHA, 1989)
(fecal coliform
bacteria/100 mL)
Sediment
MPN tubes using A-l
broth (APHA, 1989)
(fecal coliform
bacteria/100 mL)
Tissue
MPN tubes using EC
broth (APHA, 1989)
(fecal coliform
bacteria/1 00 mL)
Fecal coliform
bacteria/E.co//
Enterococci
C. perfringensb
mTEC (DuFour et al.,
1981)(E. co/^ 00 mL)
mE (Levin et al., 1975)a
(enterococci/100 mL)
MPN tubes using iron
milk (St. John et al.,
1982) (C. perfringensl
100 mL)
MPN tubes using iron
milk (St. John et al.,
1982) (C. perfringensl
100 mL)
MPN tubes using iron
milk (St. John et al.,
1982) (C. perfringensl
100mL)
aThis method is a tedious process. EPA Region 2 and the State of New Jersey have developed a
modified mE isolation technique.
h'wo laboratory techniques are available for C. perfringens: mCP by membrane filtration for water
(Bisson and Cabelli, 1979) and sediment (Emerson and Cabelli, 1982), and iron milk tube using MPN
techniques (St. John et al., 1982). The latter method is recommended (pending comparative data)
because the procedure is simpler and less costly.
samples are reincubated and examined again at the end of 48 ± 3 hr. Production of gas
or acidic growth on tubes after this time period constitutes a positive presumptive
reaction. These samples are then submitted to the confirmed phase of testing, in which
the culture is transferred to a fermentation tube containing brilliant green lactose bile
broth. The tubes are then inoculated for 48 ± 3 hr at 35 + 0.5°C. Formation of gas in
any amount of time constitutes a positive confirmed phase. The Most Probable Number
(MPN) index is then used to estimate bacterial density. This technique is commonly
used in assessing bacterial levels in shellfish.
The MPN index is an index of the number of coliform bacteria that, most probably, would
produce the results observed in the laboratory examination (APHA, 1989). Values for
the MPN index can be obtained from standard tables based on the results of the
multiple-tube fermentation technique or from Thomas's formula (APHA, 1989).
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Viruses
Detecting viruses in marine waters requires collecting a representative water sample,
concentrating viruses in the sample, and identifying and estimating abundances of these
concentrated viruses. Difficulties in detecting virus abundances include:
• Viruses are very small (20-100 nm).
• Virus concentrations in waters are spatially and temporally variable, and
typically low.
• Dissolved and suspended material in the sample interferes with accurate
virus detection.
Because virus concentrations may be very low, significant volumes of water may be
required (e.g., on the order of tens to thousands of liters). Dose-response curves are
lacking for most pathogens; however, as few as 10 to 100 bacteria are capable of
inducing disease (OTA, 1987).
Three different techniques for concentrating and enumerating viruses are presented in
Standard Methods (APHA, 1989):
• Adsorption to and elution from microporous filters;
• Aluminum hydroxide adsorption and precipitation; and
• Polyethylene glycol hydroextraction-dialysis.
The adsorption and elution technique pressure-filters viruses on microporous filters and
then elutes them from the filter in a small liquid volume. Generally, two types of filters
are available: electronegative and electropositive filters. Currently, insufficient
documentation on the efficacy of electropositive filters exists. Limitations of this
technique include:
• The adsorbent filter may be clogged by suspended material.
• Dissolved colloid material may interfere by competing with viruses for
adsorption sites on the filter.
• Viruses adsorbed to suspended material may be removed during suspended
material clean-up procedures.
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Methods for recovering solids-associated viruses are found in APHA (1989). In spite of
these limitations, the adsorption and elution technique remains the most promising
method for detecting viruses.
Aluminum hydroxide adsorption and precipitation and polyethylene glycol
hydroextraction-dialysis are used to reconcentrate viruses in proteinaceous and organic
buffer eluates. These methods may be used to concentrate viruses in waters having
high virus densities (e.g., wastewaters). They may also be used as a second-step
concentration procedure following processing of large fluid volumes through
microporous filters. However, these two techniques are impractical for primary
processing of large fluid volumes (APHA, 1989).
New Techniques
Traditional techniques lack the ability to detect apparently viable, but nonculturable,
microorganisms. Recently developed monoclonal antibody and gene probe techniques
permit the detection and enumeration of both culturable and nonculturable
microorganisms. Since these methods have the ability to detect nonculturable
organisms, they may serve as more precise techniques for monitoring microbial water
quality. Furthermore, these new techniques will allow direct monitoring of pathogens of
concern. The goal of monitoring specific pathogens, such as Salmonella, Shigella,
Giardia, or Legionella, may soon be realized.
Sample Handling
Preservation and storage of water samples can be significant sources of error. Sample
bottles must be resistant to sterilization procedures. Samples should be refrigerated
(1-4°C) during transport to a laboratory and analyzed within 6 hours of collection
(USEPA, 1978).
4.11.4 QA/QC Considerations
It is recommended that sterile distilled water be transported to the field, transferred to a
sample bottle, and processed routinely as a field blank to ensure that samples were not
contaminated during collection and transport. It is also recommended that 10 percent of
the samples be analyzed in duplicate. Furthermore, 10 percent of the samples should
be split and analyzed by two or more laboratories.
Intralaboratory and interlaboratory quality control practices should be documented, and
QC reports should be available for inspection. Further recommended quality assurance
guidelines for a microbiology laboratory are available (Inhorn, 1977; Bordner et al.,
1978) and are also discussed in APHA (1989).
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4.11.5 Statistical Design Considerations
Bacterial counts often are characterized as having a skewed distribution because of
many low values and a few high ones (APHA, 1989). Application of parametric
statistical techniques requires the assumption of symmetrical distributions such as the
normal curve. An approximate normal distribution can be obtained from positively
skewed data by converting numbers to their logarithms (APHA, 1989). Accordingly, the
preferred statistic for measuring central tendency of microbiological data is the
geometric mean.
Consideration of statistical strategies will mitigate the high costs of collecting and
processing samples. Power-cost analyses are necessary in selecting appropriate
sample/replicate number, sample location, and sampling frequency (Ferraro et al.,
1989).
4.11.6 Use of Data
The assessment of pathogen contamination should be a component of a monitoring
program where pathogens may present a risk to human health and the economic vitality
of an estuary. Monitoring of human pathogens provides information essential in relating
the temporal and spatial distribution of infectious agents in marine waters and shellfish
to the epidemiology of pathogens of concern (e.g., affected human populations,
locations, and timing of the outbreak).
Furthermore, monitoring data can be used to identify discharges that may be significant
sources of pathogens and to ensure that water quality standards are maintained.
Monitoring data can also be used to verify fate and transport, human health risk
assessment, and epidemiological modeling predictions.
4.11.7 Summary and Recommendations
Rationale
• The objective is to detect and describe spatial and temporal changes in
abundances of indicators of human health pathogens.
• Monitoring indicators of human pathogens provides information essential in
relating the presence of infectious agents in marine waters and shellfish to
the incidence of disease outbreaks and potential sources of these agents.
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Monitoring Design Considerations
• If feasible, samples of the surface microlayer should be collected separately
from samples of underlying waters. If not feasible, samples containing both
underlying and microlayer waters should be collected. It is highly
recommended that consistent types of sampling protocols be implemented to
allow for comparisons between studies.
• Analysis of tissues of sentinel organisms (e.g., mussels and oysters) confers
the following advantages:
- such organisms concentrate pathogens;
- such organisms provide a means of temporally integrating water quality
conditions; and
- such organisms may be deployed and maintained in a number of locales.
Analytical Methods Considerations
• Indicators of human pathogens should have the following characteristics:
- should be consistently found in high abundance in human fecal wastes;
- should not have significant extra-human fecal sources; and
- must provide temporally and spatially reliable and accurate appraisals of
the pathogen of concern.
• Fecal coliform bacteria densities.
- May not be a reliable indicator for predicting the risk associated with direct
exposures to pathogens.
- This measure provides a means to compare current data to historical data.
- Laboratory analyses include membrane filter and multiple-tube
fermentation techniques.
• Enterococci densities
- Densities are highly correlated with the incidence of gastrointestinal
symptoms.
- Taxonomic identifications are relatively simple and can reveal the kinds of
mammalian pollution.
- State-of-the-art genetic fingerprinting techniques can link bacteria in the
marine environment to specific sources; however, these techniques are
very costly and require further study to assess their reliability.
- Laboratory analyses include membrane filter technique.
• Virus densities
- Viruses are recognized as major etiologic agents for many outbreaks of
human illness.
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- Examination of water samples for enteric viruses is not recommended
because of methodology limitations.
- Laboratory analyses include absorption to an elution from microporous
filters, aluminum hydroxide absorption-precipitation, and polyethylene
glycol hydroextraction-dialysis.
• Monoclonal antibodies and gene probes show promise in detecting and
enumerating both culturable and nonculturable microorganisms.
QA/QC Considerations
• Sterile distilled water should be transported and transferred to sample bottles
as field blanks to assess contamination during collection and transport.
• Ten percent of the samples should be analyzed in duplicate.
• Ten percent of the samples should be split and analyzed at two or more
laboratories.
Statistical Design Considerations
« The preferred statistic for measuring central tendency of microbiological data
is the geometric mean.
Use of Data
• Provide essential information to assess threats to human health;
• Establish temporal and spatial trends in pathogen densities and identify
potential relationships between the presence of the indicator and incidence of
human illness; and
• Provide information that can be used to verify fate and transport, human
health risk, and epidemiology models.
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4.12 EFFLUENT CHARACTERIZATION
"End-of-pipe" effluent characterizations based on laboratory studies can be used to
predict biological impacts of an effluent prior to a discharge. The Technical Support
Document for Water Quality-based Toxics Control (USEPA, 1991c) presents (among
other things) an integrated approach for ensuring protection of aquatic life and human
health from impacts caused by the release of toxics to surface waters. For the
protection of aquatic life, the integrated strategy involves the use of three control
approaches: the chemical-specific control approach, the whole-effluent toxicity control
approach, and the biological criteria/bioassessment and biosurvey approach. For the
protection of human health, technical constraints do not yet allow for full reliance on an
integrated strategy, and thus primarily chemical-specific assessments and control
techniques should be employed. Table 4-30 contains a list of terms used in effluent
characterization.
Pollutant-specific criteria developed pursuant to CWA section 304(a)(1) present
scientific data and guidance on the environmental effects of pollutants that reflect the
latest EPA recommendations on acceptable limits for aquatic life and human health
protection. These limitations are generally developed from laboratory-derived,
biologically-based numeric water quality criteria adopted within a State's water quality
standards. Water quality criteria are adopted by a State for the protection of designated
uses of the receiving water. Criteria can be specific numeric limits or, in the absence of
specific numeric criteria for a chemical, biological, or physical parameter, can be based
on narrative criteria. EPA publishes chemical-specific water quality criteria that may
provide the basis on which States adopt their criteria (USEPA, 1986-1988). EPA has
published 25 chemical-specific saltwater aquatic life criteria (acute and chronic).
When an effluent's constituents are not completely known or when a complex mixture of
potentially additive, antagonistic, or synergistic toxic pollutants is discharged, a
whole-effluent toxicity limitation can be. implemented. The whole-effluent toxicity
limitation can also be appropriate when more than one discharger is located in a specific
area and the potential exists for effluent mixing and additive toxic effects, and when a
pollutant-specific evaluation is impractical because of a lack of information about the
toxic effects of a pollutant. Toxicity evaluations such as these can be used as part of
403 monitoring.
4.12.1 Rationale
Effluent characterization can be used in the CWA section 403 point source discharge
program to address the following evaluation criteria:
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Table 4-30. List of Terms Used with Effluent Characterization
Acute
Acute-to-chronic ratio
(ACR)
Chronic
Composite sample
Effective concentration
Grab samples
A stimulus severe enough to rapidly induce an effect; in aquatic toxicity tests,
an effect observed in 96 hours or less typically is considered acute. When
referring to aquatic toxicology or human health, an acute effect is not always
measured in terms of lethality.
The ratio of the acute toxicity of an effluent or a toxicant to its chronic toxicity.
ACR is used as a factor for estimating chronic toxicity on the basis of acute
toxicity data, or the reverse.
A stimulus that lingers or continues for a relatively long period of time, often
one-tenth of the life span or more. Chronic should be considered a relative
term depending on the life span of an organism. The measurement of a
chonic effect can be reduced growth, reduced reproduction, etc., in addition
to lethality.
A single effluent sample collected over a 24-hour period, on which only one
toxicity test is performed.
A point estimate of the toxicant concentration that would cause an
observable adverse effect (such as death, immobilization, or serious
incapacitation) in a given percentage of the test organisms.
Samples collected over a very short period of time and on a relatively
infrequent basis. A separate toxicity test must be performed on each grab
sample.
Lowest-Observed-Adverse- The lowest concentration of an effluent or toxicant that results in statistically
Effect Level (LOAEL) significant adverse health effects as observed in chronic or subchronic
human epidemiology studies or animal exposure.
No-Observed-Adverse-
Effect Level (NOAEL)
No-Observed-Effect
Concentration (NOEC)
Toxicity characterization
Whole-effluent toxicity
A tested dose of an effluent or a toxicant below which no adverse biological
effects are observed, as identified from chronic or subchronic human
epidemiology studies or animal exposure studies.
The highest tested concentration of an effluent or a toxicant at which no
adverse effects are observed on the aquatic test organisms at a specific time
of observation. Determined using hypothesis testing.
A determination of the specific chemicals responsible for effluent toxicity.
The aggregate toxic effect of an effluent (usually measured directly with a
toxicity test).
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• The quantities, composition, and potential for bioaccumulation or persistence
of the pollutants to be discharged;
• The potential impact on human health through direct and indirect pathways; and
• Marine water quality criteria developed pursuant to section 304(a)(1).
Assessment of these evaluation criteria can be made through whole-effluent toxicity
testing (acute and chronic) and analysis of the chemical composition of the effluent.
This assessment would demonstrate potential impacts from the effluent in question.
4.12.2 Monitoring Design Considerations
Determination of whether an effluent sample is typical of the wastewater may require the
collection of a large number of samples. Further, what constitutes a "representative"
sample is a function of the parameter of concern. Guidelines for determining the
number and frequency of samples required to represent effluent quality are contained in
the Handbook for Sampling and Sample Preservation of Water and Wastewater
(USEPA, 1982b).
Both quantitative (change in concentration) and qualitative (change in toxicants)
variability commonly occur in effluents. Changes in effluent toxicity are the result of
varying concentrations of individual toxicants, different toxicants, changing in-stream
water quality characteristics (affecting compound toxicity), and analytical and
toxicological error. Conventional parameters, BOD, TSS, and other pollutants limited in
the facility's permit will provide an indication of the operational status of the treatment
system on the day of sampling. This information may be used in a determination as to
whether the system is operating properly, requires repair, or needs to be upgraded. For
industrial discharges, information on production levels and types of operating processes
may be helpful in determining the required magnitude of the monitoring program.
4.12.3 Analytical Methods Considerations
Sampling
The choice of grab or composite samples will depend on the specific discharge situation
(e.g., plant retention time) and the objectives of the test. In toxicity characterization
testing, samples that are very different from one another give results that are difficult to
interpret. However, composite sampling has an averaging effect, which tends to dilute
toxicity peaks and may thus provide misleading results when testing for acute toxicity.
Composited samples, therefore, are more appropriate for chronic toxicity tests where
peak toxicity of short duration is of less concern. If the toxicity of the effluent is variable,
grab samples collected during peaks of effluent toxicity provide a measure of maximum
effect. However, while grab sampling may provide information on maximum effluent
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toxicity, it can be difficult to schedule sampling to coincide with peaks in toxicity.
Aeration during collection and transfer of effluents should be minimized to reduce the
loss of volatile chemicals.
Chemical-Specific Analysis
Pollutant-specific testing for comparison to water quality criteria should be done
following standard methods. Choice of a particular method may vary depending on the
precision and accuracy required and the resources available for the study. A review of
water chemistry analytical methods is presented in the section in this report on water
quality and in Standard Methods for the Examination of Water and Wastewater (APHA,
1989). However, it should be noted that pollutant concentrations are considerably
higher in effluent streams and the precision required for adequate measurement is
lower.
Whole-Effluent Toxicity Analysis
Whole-effluent toxicity testing can be somewhat more complex than a pollutant-specific
approach. Upon arrival of the sample in the laboratory, temperature, pH, hardness, and
conductivity should be measured. Total residual chlorine, total ammonia, alkalinity,
dissolved oxygen (DO), and organic carbon measurements may also be appropriate.
These measurements can provide necessary information should the toxicity of the
effluent change over time.
EPA recommends that for whole-effluent toxicity data generation the process should be
divided into three basic steps: initial dilution determination, toxicity testing procedures,
and triggers for permit limit development. The dilution determination is an estimate of
the effluent dilution at the edge of the mixing zone and should take any applicable State
mixing zone requirements into consideration.
Whole-effluent toxicity testing should entail both acute and chronic testing. An acute
toxicity test is defined as a test of usually less than 96 hours in duration in which lethality
is the measured endpoint. A chronic toxicity test is defined as a long-term test in which
sublethal effects, such as fertilization, growth, and reproduction are usually measured,
although in highly toxic effluents lethality may also result. Traditionally, chronic tests are
full-life-cycle tests or at least 30-day tests. However, the duration of most of the EPA
chronic toxicity tests has been shortened to 7 days by focusing on the most sensitive
life-cycle stages. For this reason, the EPA chronic toxicity tests are called short-term
chronic tests. Whole-effluent toxicity limits can be inserted into permits as a means of
controlling pollutants when chemical-specific criteria have not been developed or
synergistic effects are found to occur.
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For whole-effluent toxicity testing, EPA recommends that three species (a vertebrate, an
invertebrate, and a plant) be tested quarterly for a minimum of 1 year. This may be
adjusted to require testing of only the most sensitive species. Conducting tests quarterly
for 1 year is recommended to adequately assess the variability of toxicity observed in
effluents. The use of three of the five commonly used marine organisms—inland
silverside, sheephead minnow, mysid shrimp, Champia, and sea urchin—has generally
been sufficient to measure the toxicity of any effluent for the purposes of projecting
effluent toxicity impact and making regulatory decisions. Specific test methodologies
can be found in a variety of EPA manuals, including Methods for Aquatic Toxicity
Identification Evaluations - Phases I, II, and III (USEPA, 1988a, b, c), Methods for
Measuring the Acute Toxicity of Effluents to Freshwater and Marine Organisms
(USEPA, 1985c), Short-Term Methods for Estimating the Chronic Toxicity of Effluents
and Receiving Waters to Marine and Estuarine Organisms (USEPA, 1988e), and
Biomonitoring for Control of Toxicity in Effluent Discharges to the Marine Environment
(USEPA, 1989b). Numerous other sources of toxicity testing methodologies exist (see
reference list) and may be more appropriate for the effluent of interest.
Biological Criteria/Bioassessment
To fully protect aquatic habitats, water quality criteria should address biological integrity.
Biocriteria should be used as a supplement to existing chemical-specific criteria and as
criteria where such chemical-specific criteria have not been established. Biocriteria are
numerical values or narrative expressions that describe the reference biological integrity
of aquatic communities inhabiting unimpaired waters. The biological communities in
these waters represent the best attainable conditions (USEPA, 1991c).
Resident biota integrate multiple impacts over time and can detect impairment from
known and unknown causes. Biocriteria can be used to verify improvement in water
quality in response to regulatory efforts and detect continuing degradation of waters.
Numeric criteria can provide effective monitoring criteria for inclusion in permits
(USEPA, 1991).
The assessment of biological integrity should include measures of the structure and
function of an aquatic community of species within a specified habitat. Expert
knowledge of the system is required for the selection of appropriate biological
components and measurements indices. See Section 4.5 for further discussion of
benthic community structure.
4.12.4 QA/QC Considerations
Basic sampling and laboratory quality assurance and quality control (QA/QC)
procedures for pollutant-specific characterization are addressed in the section on water
chemistry. QC problems specific to effluent characterization include variability in effluent
concentrations and toxicity response of test organisms, and changes in effluent toxicity
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over time. A quality assurance project plan (QAPj'P) should be developed and adhered
to. The QAPjP should include quality verification, which entails a demonstration that the
proposed study plan was followed as detailed and that work carried out was properly
documented. The QAPjP should increase communication between clients, program
planners, field and laboratory personnel, and data analysts. The QAPj'P must make
clear the specific responsibilities of each individual. The QC procedures involve
standardized guidelines, such as the number of samples to be taken and the mode of
collection, standard operating procedures for analyses, and spiking protocols. QC
protocols for standard toxicity test methods are described in USEPA (1985c).
4.12.5 Statistical Design Considerations
Statistical design of an effluent characterization project must account for variability in
effluent and toxicity testing response. However, the design of any given experiment
must be weighed against the importance of the data and decisions to be based on the
data. The critical nature of certain data will demand stringent controls, while statistical
rigor (and associated costs) can be lessened in other experiments having less impact
(e.g., initial toxicity testing).
The choice of a statistical method to analyze toxicity test data and the interpretation of
the results of the analysis of the data from toxicity tests can be problematic because of
the inherent variability and sometimes unavoidable anomalies in biological data. The
assistance of a trained statistician is recommended before selecting the method of
analysis and interpretation of the results. The data should be plotted to help detect
problems and unsuspected trends or patterns in the responses and to aid in
interpretation of the results. The analysis of the data is dependent upon the number of
replicates and the distribution and homogeneity of the data. A variety of statistical
methods for the analysis of toxicity data (including Dunnett's procedure, Bonferroni's
t-test, Steel's many-one rank test, the Wilcoson Rank Sum test, and the Probit analysis)
are presented in USEPA (1988e).
4.12.6 Use of Data
Pollutant-specific effluent characterization can be used for comparison to State water
quality standards. A chronic criterion is a level at which aquatic organisms and their
uses should not be affected unacceptably if the 4-day average concentration of the
pollutant is not exceeded more than once every 3 years on average. An acute criterion
is a limit at which the pollutant concentration should not be exceeded more than once
every 3 years on average. Care should be taken when applying these criteria to assess
potential impacts on locally important species that are very sensitive.
Whole-effluent toxicity testing data (both acute and chronic) may be used to screen an
effluent for potential unreasonable degradation of the environment. Whole-effluent
toxicity is a useful parameter for assessing and protecting against impacts on water
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quality and designated uses caused by the aggregate toxic effect of the discharge.
Thus, toxicity itself is used as the effluent parameter, and the toxicants creating the
toxicity need not be specifically identified or controlled unless a Toxicity Reduction
Evaluation (TRE) is required to ensure compliance with a toxicity limit in the permit.
If a threshold of toxicity is desired, then the results of whole-effluent toxicity testing may
be used to derive the No-Observed-Effect Concentration (NOEC) and the
Lowest-Observed-Effect Concentration (LOEC). Whole-effluent toxicity testing can also
produce a concentration-dependent result of some amount of adverse effect called the
Effective Concentration (EC). For example, ECso is the effluent concentration that
would affect 50 percent of the organisms tested. Interpretation of EC values, therefore,
requires the judgment of a toxicologist.
4.12.7 Summary and Recommendations
Rationale
• Effluent characterization may be used to assess potential impacts, both acute
and chronic, of the effluent on the surrounding biological communities without
interference from other sources.
Monitoring Design Considerations
• Effluent variability, including potential changes in contaminants and
contaminant concentrations, should be considered when designing a
monitoring plan.
Analytical Methods Considerations
• Grab or composite samples should be used depending on the objectives and
type of test performed.
• Chemical-specific testing can be performed to characterize toxicity and
ensure compliance with water-quality standards.
• Both acute and chronic tests can be used to assess whole-effluent toxicity.
QA/QC Considerations
• See section on water chemistry.
• QAPjP should be developed, adhered to, and documented.
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Statistical Design Considerations
• Effluent variability and toxicity testing response must be accounted for in the
monitoring plan.
Use of Data
• Chronic and acute toxicity levels can be set in permits.
• Toxicity Reduction Evaluations (TREs) can be required.
• Threshold and effective concentrations can be established to predict
ecological risk.
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4.13 MESOCOSMS AND MICROCOSMS
The use of mesocosms and microcosms for the assessment of ecological impacts from
marine dischargers is considered to be premature. This monitoring approach is the
focus of research and holds promise for future use. It has been used to some extent by
EPA's Office of Pesticides to predict impacts from the use of pesticides.
Microcosms and mesocosms, on a small to medium size scale, respectively, attempt to
simulate the complex nature of natural environments. An effort is made to determine the
response of an ecosystem to perturbation through an accurate simulation of the
physical, chemical, and biological characteristics of an ecosystem and the biological
interactions of the organisms in the ecosystem. While there is no absolute separation
between a mesocosm and a microcosm, in practice most mesocosms are outdoors
where temperature and light intensity are ambient for the parent community. Most
mesocosm researchers try to realistically scale for the factors they think are most
important in controlling ecological relationships. Most microcosms are incubated in the
laboratory and are subject to greater environmental control. Microcosms are usually
smaller, and more replicates are used.
4.13.1 Rationale
Of the 10 guidelines used to determine unreasonable degradation or no irreparable
harm, several can be addressed through the use of mesocosms and microcosms
including:
• The potential transport of such pollutants by biological, physical or chemical
processes and
• The composition and vulnerability of potentially exposed biological
communities.
Mesocosms and microcosms can be used to gain a realistic characterization of
ecosystem-level impacts of marine discharges subject to section 403. With mesocosms
and microcosms, the responses of complete ecosystems to pollutants can be
investigated. Other advantages include the possibility of experimentation, at near
natural conditions, with addition of isotopic tracers to an ecosystem and the opportunity
to test numerical transport models under near natural conditions.
If insufficient information exists to determine whether a discharge will result in
unreasonable degradation of the marine environment, EPA must make a determination
as to whether a discharge will cause irreparable harm during the period in which
monitoring is undertaken. The use of microcosms allows for the incorporation of
ecosystem recovery (a critical factor in making a determination of irreparable harm) as
an endpoint for future ecological risk assessment (Perez et al., 1990).
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4.13.2 Monitoring Design Considerations
A microcosm should be designed to represent a miniature ecosystem that responds
quickly to perturbation. The emphasis is on ecosystem-level properties, although
species-level attributes such as fecundity and mortality can also be monitored if desired.
Most generic microcosms do not simulate a site-specific natural system, but function as
very generalized simulations of a large class of ecosystems. However, the use of
undisturbed, natural aquatic, and benthic communities (through the use of a sediment
core) in a single system allows for the simulation of a natural system (USEPA, 1983b).
The design of the microcosm allows the system to be defined in terms of its physical and
temporal boundaries, its light and temperature regime, its water composition, the
turbulence and turnover rate, the ratio of benthic surface area to seawater volume, the
sediment characteristics, and the water flow rate over the sediment surface.
The results of microcosm or mesocosm experiments are measured by the ecological
effects on the biota of the system. When selecting biota to include in the test, two
factors should be considered, both of which involve selection of the benthic component
of the system. If the natural system has more than one distinct benthic community, then
those organisms which are directly linked to human consumption or those which are
important to the economics and aesthetics of the area should be considered for
experimental use. Moreover, if some of the benthic communities contain species known
to be more sensitive to environmental contaminants than others, these communities
should also be considered.
4.13.3 Analytical Methods Considerations
Microcosms for which ecosystem effects protocols have been developed are small,
static, open ecosystems. Kenneth Perez and other researchers at the EPA
Environmental Research Laboratory in Narragansett, Rhode Island, developed an
experimental marine microcosm test protocol that employs a time frame of 30 days
(USEPA, 1983b). The methodology was published in the Federal Register (volume 52,
number 187, pages 36352-36360) for use in developing site-specific data on the
chemical fate and ecological effects of chemical substances subject to regulation under
the Toxic Substances Control Act (TSCA) and could be adapted to the 403 ocean
discharge program. This protocol attempts to couple undisturbed pelagic and benthic
communities within a single system, the physical and chemical conditions of which are
equated to those in the natural system being simulated. The protocol and support
document (USEPA, 1990e) describes the steps necessary to develop the experimental
microcosm (tanks, paddles, benthic cylinders, pumps, and pump air supply) and support
equipment (room, water bath, light and turbulence fixtures, test compartments, and an
air evacuation system). A diver-collected sediment core is used for the benthic
subsystem, and the test water is collected by a nondestructive method and exchanged
at least three times a week, coinciding with biological and chemical sample collection.
Water flow, turbulence, light intensity, temperature, and simulated tidal flow are all
controlled to simulate the natural system for the 30-day time period.
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Methods
Other researchers contend that to study ecosystem-level effects it is essential that the
microcosms function (1) as homeostatic, self-sustaining ecosystems capable of existing
through at least a year and (2) independent of outside subsidies except for light and
replacement of evaporated water. Each microcosm is considered a unique ecosystem
and intraexperiment replication among microcosms is enhanced by using defined
chemical media and standard physical conditions (light, temperature, day length).
Inoculation with well-developed interactive couplings of the organisms is essential to
achieve these goals. Selection of the exact species to be used should be based on the
specific objectives of the study.
During microcosm studies, effluent to be tested should be added on a gradient. A
typical experimental design might consist of six replicates in each of four treatment
groups: a control, a single addition of a low concentration, a single addition of a high
concentration, and repeated additions of low concentrations. Example protocol steps
and the variables monitored for this type of generic microcosm are described in Taub
(1984).
The physical structure of a mesocosm system determines the ability of water,
organisms, and pollutants to move from compartment to compartment within the system.
Coastal ecosystems and the fate and effects of pollutants within these systems are
strongly affected by vertical mixing and the interaction of planktonic and benthic
processes. Donaghay (1984) classified mesocosm systems by their degree of
planktonic/benthic coupling into three categories: single well-mixed benthic coupled,
single stratified, and totally benthic decoupled. Single well-mixed benthic coupled
systems are composed of a well-mixed water column in direct contact with sediment and
benthos. This system has been used extensively at the Marine Ecosystem Research
Laboratory (MERL) with very good results validating field data; defining loadings, fates,
and effects; and manipulating natural systems to define underlying processes. The
single stratified system design involves well-mixed top and bottom layers separated by
an unmixed thermocline with only the bottom layer in contact with the sediment. The
system is used to study systems where stratification is important. The single system
without benthic coupling is intended to model fates and effects in stratified deep water
systems without benthic coupling.
4.13.4 QA/QC Considerations
In some cases the accuracy of the results of mesocosms and microcosms may be
limited by a number of factors. The size of the systems most often excludes
macrofauna such as fish and large crustaceans. The size of a microcosm can also
affect the mixing rates of surficial sediments. An increase in the size of a microcosm will
result in an increase in exposure of sediment particles to the overlying water column.
Results from smaller-sized systems have underestimated potential risks of complex
effluent (Perez et al., in press; Pontasch et al., 1989); however, attempts are being
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Section 403 Procedural and Monitoring Guidance
made to develop "scaling laws" to describe this relationship. These systems are also
limited to photo-stable test substances because the light intensity may be higher than
that which would occur naturally.
4.13.5 Statistical Design Considerations
The results of the microcosm and mesocosm experiments assist in determining the fates
and ecological effects of contaminants within a marine system. The pelagic biota are
characterized by the number and species composition of phytoplankton, zooplankton,
and transient larval forms. The benthic community is characterized by the structural
composition (see Section 4.5, Benthic Community Structure). The experimental marine
test protocol (USEPA, 1990e) recommended test design and statistical analysis allow for
the independent assessment of the solvent carrier (if used) and the effluent for all
variables measured. Also, the differences in the biotic response between the control
microcosms and the natural system will provide a measure of the validity of the test
response. A multivariate analysis of variance, followed by univariate analysis, and
regression techniques are recommended for the analysis of all data.
Statistical design considerations of mesocosms must consider that the scientific and
financial resources for mesocosm studies are limited. Therefore, efforts must be made
to maximize the information gained and the statistical rigor of the analysis while
minimizing the number of systems analyzed. Four different experimental designs are
presented by Donaghay (1984): single system, paired system, replicated, and single
gradient. Each design has advantages and disadvantages, and each is appropriate for
specific purposes.
4.13.6 Use of Data
Mesocosm research and microcosm research serve somewhat different purposes. If a
researcher wants to have maximum confidence in extrapolating back to a specific,
large-scale environment, the mesocosm provides greater realism in that it allows more
large-scale processes to be included in the system. If simplification is the goal—either
to semi-isolate certain components to determine their importance in the degradation of
pollutants or to provide more easily analyzed ecosystems for test purposes—then
microcosms are more appropriate.
The use of microcosms and/or mesocosms in an ocean discharge criteria evaluation as
mandated under section 403 would be most appropriate for general permits issued for a
whole category of discharges (e.g., regional oil and gas exploration/production activities)
mainly because of the higher cost associated with implementing a series of such
experiments. However, the use of microcosms and mesocosms can be less expensive
than implementing field experiments and can provide quantitative estimates of
processes and responses that can only be assessed qualitatively with field experiments.
In addition, Perez and Morrison (1985) showed that the monetary costs of environmental
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Methods
assessments of a single chemical using a single microcosm system versus a series of
simple bioassay and physicochemical test systems are, at the worst case, effectively
equivalent. Assessments of complex effluents performed using microcosms could result
in more accurate results with a potential cost savings.
The use of microcosms and/or mesocosms could add significant benefits to an
ecological risk assessment of an ocean discharge. Microcosms and/or mesocosms can
be used to estimate ecological effects, chemical fate, transport mechanisms,
bioaccumulation, and ecological risk. Where multiple discharges occur in one area,
microcosms and/or mesocosms could be used to assess the risk associated with
individual discharges or combinations of discharges. Estimating ecosystem recovery
through the use of mesocosms in ecological risk assessments is a recent application
(Perez et al., 1990) and could provide important information concerning the occurrence
of irreparable harm.
4.13.7 Summary and Recommendations
Rationale
• Mesocosms and microcosms can be used to gain a realistic characterization
of ecosystem-level impacts of marine discharges subject to section 403(c).
• The transport and fate of pollutants to the complete ecosystems can be
investigated in near natural conditions.
Monitoring Design Considerations
• A microcosm should be designed to represent a miniature ecosystem that
responds quickly to perturbation.
• Most generic microcosms do not simulate a site-specific natural system, but
function as very generalized simulations of a large class of ecosystems.
• The design of the microcosm allows the system to be defined in terms of its
physical and temporal boundaries, its light and temperature regime, its water
composition, the turbulence and turnover rate, the ratio of benthic surface
area to seawater volume, the sediment characteristics, and the water flow
rate over the sediment surface.
• The results of microcosm or mesocosm experiments are measured by the
ecological effects on the biota of the system.
• Microcosm studies have been designed for 30 days to more than a year in
duration.
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Section 403 Procedural and Monitoring Guidance
Analytical Method
• During microcosm studies, effluent to be tested should be added on a
gradient. A typical experimental design might consist of six replicates in each
of four treatment groups: a control, a single addition of a low concentration, a
single addition of a high concentration, and repeated additions of low
concentrations.
• Mesocosm systems can be classified by their degree of planktonic/benthic
coupling into single, well-mixed benthic coupled; single stratified; and totally
benthic decoupled systems.
QA/QC Analysis
• The results of microcosm and mesocosm studies are limited by the size of
the system.
• Scaling laws may be used to translate from microcosms to full-scale
ecosystems.
Statistical Design Considerations
• A multivariate analysis of variance, followed by univariate analysis, and
regression techniques are recommended for the analysis of all data.
• Efforts must be made to maximize the information gained and the statistical
rigor of the analysis while minimizing the number of systems analyzed.
Use of Data
• Data can be used to estimate ecological effects, chemical fate, transport
mechanisms, bioaccumulation, and ecological risks.
• Data can also be used to assess impacts of individual or multiple discharges
and to estimate recovery time.
• Mesocosms are most appropriate for use when maximum confidence in
extrapolating back to a specific, large-scale environment is desired.
• Microcosms are most appropriate for use when simplification is the goal,
either to isolate certain components to determine their importance or to
provide more easily analyzed ecosystems for test purposes.
• Microcosms and rnesocosms may be most appropriate for use for general
permits issued for a whole category of discharges.
• There can be cost advantages to using microcosms and rnesocosms under
certain circumstances.
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APPENDIX A:
MONITORING METHODS REFERENCES
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Appenctfx A
PHYSICAL CHARACTERISTICS
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Appendix A
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Appendix A
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WATER CHEMISTRY
APHA. 1989. American Public Health Association, American Water Works Association,
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Appendix A
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Environmental Protection Agency, Office of Marine and Estuarine Protection,
Washington, DC.
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A-7
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Appendix A
Ward, B.C., and J.C. Loftis. 1986. Establishing statistical design criteria for water quality
monitoring systems, review and synthesis. Water Res. Bull. 22(5): 759-767.
SEDIMENT CHEMISTRY
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A-8
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Appendix A
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A-9
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Appendix A
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A-10
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Appendix A
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A-41
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Appendix A
State of Maryland Department of Natural Resources. Undated. Environmental
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Appendix A
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Appendix A
Ladd, J.M., S.P. Hayes, M. Martin, M.D. Stephenson, S.L. Coale, J. Linfield, and
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A-45
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Appendix A
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USEPA. 1990. Computerized Risk and Bioaccumulation System (Version 1.0).
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A-47
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Appendix A
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A-48
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Appendix A
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A-49
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Appendix A
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Appendix A
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A-53
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Appendix A
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Appendix A
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EFFLUENT CHARACTERIZATION
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Appendix A
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PJ. Shubat. 1986. Comparative toxicity of methanol and
N.N-dimethylformamide to freshwater fish and invertebrates. Bull. Environ.
Contam. Toxicol. 37(4) :6 1 5-621 .
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wet oxidation. J. Water Pollut. Control Fed. 52(8): 217-2130.
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program: Test sensitivity, repeatability, and relevance to receiving water toxicity. Env.
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chemical equilibria in natural waters. John Wiley and Sons, Inc., New York, NY.
U.S. Department of Health, Education and Welfare. 1977. Carcinogens - working with
carcinogens. Publication no. 77-206. Public Health Service, Center for Disease
Control, National Institute of Occupational Safety and Health.
US EPA. 1982. Handbook for sampling and sample preservation of water and
wastewater. EPA 600/4-82-01 9. U.S. Environmental Protection Agency, Environmental
Monitoring and Support Laboratory, Cincinnati, OH.
USEPA. 1982. Test methods - Technical additions to methods for chemical analysis of
water and wastes. EPA 600/4-82-055. U.S. Environmental Protection Agency, Office of
Research and Development, Cincinnati, OH. December.
USEPA. 1982. Water quality assessment: A screening procedure toxic and
conventional pollutants. Parts 1 and 2. EPA 600/6-82-004. U.S. Environmental
Protection Agency, Office of Research and Development, Athens, GA.
A-56
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Appendix A
USEPA. 1983. The treatability manual. Vol. IV. EPA 600/2-82-001. U.S.
Environmental Protection Agency, Office of Research and Development. GPO,
Washington, DC.
USEPA. 1984. Effluent and ambient toxicity testing and instream community response
on the Ottawa River, Lima, Ohio. EPA 600/2-84-080. Permits Division, Washington,
DC, U.S. Environmental Protection Agency, Office of Research and Development,
Duluth, MN.
USEPA. 1984. CETIS: Complex Effluent Toxicity Information System. Data encoding
guidelines and procedures. EPA 600/8-84-029. U.S. Environmental Protection Agency,
Office of Research and Development, Duluth, MN.
USEPA. 1984. CETIS: Complex Effluent Toxicity Information System. CETIS retrieval
system user's manual. EPA 600/8-84-030. U.S. Environmental.Protection Agency,
Office of Research and Development, Duluth, MN.
USEPA. 1984. Development of water quality based permit limitations for toxic
pollutants; National policy. U.S. Environmental Protection Agency. Fed. Regist, March
9,1984,49(48).
USEPA. 1984. Technical guidance manual for performing wasteload allocations, Book
III estuaries. U.S. Environmental Protection Agency, Office of Water Regulations and
Standards, Washington, DC.
USEPA. 1985. Methods for measuring the acute toxicity of effluents freshwater and
marine organisms. EPA 600/4-85-013. U.S. Environmental Protection Agency,
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
USEPA. 1985. Short-term methods for estimating the chronic toxicity of effluents and
receiving waters to freshwater organisms. EPA 600/4-85-014. U.S. Environmental
Protection Agency, Cincinnati, OH.
USEPA. 1986-1988. Quality criteria for water. EPA 440/5-86-001. U.S. Environmental
Protection Agency, Office of Water Regulations and Standards, Washington, DC.
USEPA. 1988. Methods for aquatic toxicity identification evaluations: Phase I toxicity
characterization procedures. Draft EPA research series report. EPA 600/3-88-034.
U.S. Environmental Protection Agency, Environmental Research Laboratory, Duluth,
MN.
A-57
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Appendix A
USEPA. 1988. Methods for aquatic toxicity identification evaluations: Phase II toxicity
identification procedures. Draft EPA research series report. EPA 600/3-88-035. U.S.
Environmental Protection Agency, Environmental Research Laboratory, Duluth, MM.
USEPA. 1988. Methods for aquatic toxicity identification evaluations: Phase III toxicity
confirmation procedures. Draft phase III toxicity series report. EPA/600/3-88-036. U.S.
Environmental Protection Agency, Environmental Research Laboratory, Duluth, MM.
USEPA. 1988. Short-term methods for estimating the chronic toxicity of effluents and
receiving waters to marine and estuarine organisms. EPA 600/4-87-02. U.S.
Environmental Protection Agency, Office of Research and Development, Cincinnati, OH.
USEPA. 1988. Draft generalized methodology for conducting industrial toxicity
reduction evaluations (TREs). Draft EPA Research Series Report. U.S. Environmental
Protection Agency, Water Engineering Research Laboratory, Cincinnati, OH.
USEPA. 1989. Toxicity reduction evaluation protocol for municipal wastewater
treatment plants. EPA 600/2-88-062. U.S. Environmental Protection Agency, Water
Engineering Research Laboratory, Cincinnati, OH.
USEPA. 1989. Short-term methods for estimating the chronic toxicity of effluents and
receiving waters to freshwater organisms. EPA 600/4-89-001. U.S. Environmental
Protection Agency, Water Engineering Research Laboratory, Cincinnati, OH.
USEPA. 1989. Biomonitoring for control of toxicity in effluent discharges to the marine
environment. EPA 625/8-89-015. U.S. Environmental Protection Agency, Office of
Research and Development, Center for Environmental Research Information.
USEPA. 1990. Permit writer's guide for marine and estuarine discharges. Draft. U.S.
Environmental Protection Agency, Office of Water Enforcement and Permits,
Washington, DC.
USEPA. 1990. Assessment and control of bioconcentratable contaminants in surface
waters. Draft. U.S. Environmental Protection Agency, Office of Water Enforcement and
Permits, Washington, DC.
USEPA. 1991. Technical support document for water quality-based toxics control.
EPA 505/2-90-001. U.S. Environmental Protection Agency, Office of Water
Enforcement and Permits, Office of Water Regulations and Standards, Washington, DC.
Walters, C.I., and C.W. Jameson.
Butterworth Publ., Woburn, MA.
1984. Health and safety for toxicity testing.
A-58
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Appendix A
MESOCOSMS AND MICROCOSMS
Davey, E.W., K.T. Perez, A.E. Soper, N.F. Lackie, G.E. Morrison, R.L Johnson, and J.
F. Heltsche. In press. Significance of the surface micro-layer to the environmental fate
of di(2-ethylhexyl) phthalate predicted from marine microcosms. U.S. Environmental
Protection Agency, Environmental Research Laboratory, Ecosystems Effects Branch,
Narragansett, Rl.
Donaghay, P.L. 1984. Utility of mesocosms to assess marine pollution. In Concepts in
marine pollution measurements, ed. H.H. White, pp. 589-620. Maryland Sea Grant
College, College Park, MD.
Dwyer, R.L., and K.T. Perez. 1983. An experimental examination of ecosystem
linearization. The American Naturalist 121 (3):305-323.
Grassle, J.P., and J.F. Grassle. 1984. The utility of studying the effects of pollutants on
single species populations in benthos of mesocosms and coastal ecosystems. In
Concepts in marine pollution measurements, ed. H.H. White, pp. 621-642. Maryland
Sea Grant College, College Park, MD.
Grice, G.W. 1984. Use of enclosures in studying stress on plankton communities. In
Concepts in marine pollution measurements, ed. H.H. White, pp. 563-575. Maryland
Sea Grant College, College Park, MD.
Leffler, J.W. 1984. The use of self-selected, generic aquatic microcosms for pollution
effects assessment. In Concepts in marine pollution measurements, ed. H. White, pp.
139-158. Maryland Sea Grant College, College Park, MD.
Oviatt, C.A. 1984. Ecology as an experimental science and management tool. In
Concepts in marine pollution measurements, ed. H.H. White, pp. 539-548 Maryland
Sea Grant College, College Park, MD.
Perez, K.T., E.W. Davey, N.F. Lackie, G.E. Morrison, P.G. Murphy, A.E. Soper, and
D.L. Winslow. 1984. Environmental assessment of phthalate ester, di(2-ethylhexyl)
phthalate (DEHP), derived from a marine microcosm. Special Technical Publication
802. American Society for Testing and Materials (ASTM), Philadelphia, PA.
Perez, K.T., and G.E. Morrison. 1985. Environmental assessments from simple test
systems and a microcosm: Comparisons of monetary costs. In Multispecies toxicity
testing, ed. J. Cairns, pp. 89-95.
A-59
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Appendix A
Perez, K.T., E.W. Davey, G.E. Morrison, J.A. Cardin, N.F. Lackie, A.E. Soper,
R.J. Blasco, C. Bearce, R.L. Johnson, and S. Marino. 1989. Influence of organic
matter and industrial contaminants in sewage effluent on marine ecosystems.
ERLN Publication. U.S. Environmental Protection Agency, Environmental
Research Laboratory, Ecosystems Effects Branch, Narragansett, Rl.
Perez, K.T., E.W. Davey, J. Heltsche, J.A. Cardin, N.F. Lackie, R.L. Johnson,
R.J. Blasco, A.E. Soper, and E. Read. 1990. Recovery of Narragansett Bay, Rl:
A feasibility study. ERLN Contribution No. 1148. U.S. Environmental Protection
Agency, Environmental Research Laboratory, Ecosystems Effects Branch,
Narragansett, Rl.
Perez, K.T., G.E. Morrison, E.W. Davey, N.F. Lackie, A.E. Soper, R.J. Blasco,
D.L. Winslow, R.L. Johnson, P.G. Murphy, J.F. Heltsche. In press. Influence of
size on the fate and ecological effects of the pesticide kepone in a physical
simulation model. U.S. Environmental Protection Agency, Environmental
Research Laboratory, Ecosystems Effects Branch, Narragansett, Rl.
Pilson, M.E.Q. 1984. Should we know the fates of pollutants. In Concepts in marine
pollution measurements, ed. H.H. White, pp. 575-588. Maryland Sea Grant College,
College Park, MD.
Pontasch, K.W., B.R. Niederlehner, and J. Cairns, Jr. 1989. Comparisons of
single-species microcosm and field responses to a complex effluent. Environ. Tox. and
Chem. 8:521 -532.
Pritchard, P.H., and A.W. Bourquin. 1984. A perspective on the role of microcosms in
environmental fates and effects assessments. In Concepts in marine pollution
measurements, ed. H.H. White, pp. 117-138. Maryland Sea Grant College, College
Park, MD.
Santschi, P.H., U. Nyffeler, R. Anderson, and S. Schiff. 1984. The enclosure as a tool
for the assessment of transport and effects of pollutants in lakes. In Concepts in marine
pollution measurements, ed. H.H. White, pp. 549-562. Maryland Sea Grant College,
College Park, MD.
Taub, F.B. 1984. Introduction to laboratory microcosms. In Concepts in marine
pollution measurements, ed. H.H. White, pp. 113-116. Maryland Sea Grant College,
College Park, MD.
A-60
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Appendix A
Taub, F.B. 1984. Measurement of pollution in standardized aquatic microcosms. In
Concepts in marine pollution measurements, ed. H.H. White, pp. 159-192. Maryland
Sea Grant College, College Park, MD.
USEPA. 1983. Project summary: Experimental marine microcosm test protocol and
support document. EPA-600/S3-83-055. U.S. Environmental Protection Agency,
Environmental Research Laboratory, Narragansett, Rl.
USEPA. 1987.
36352-36360.
Site-specific aquatic microcosm test. Fed. Regist. 52(187):
USEPA. 1990. Experimental marine microcosm test protocol and support document.
Revised. U.S. Environmental Protection Agency, Environmental Research Laboratory,
Ecosystems Effects Branch Narragansett, Rl.
A-61
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APPENDIX B:
OCEAN DISCHARGE CRITERIA
PUBLISHED AT FR Vol. 45, No, 194, 65942-65954
OCTOBERS, 1980
image:
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J35942
Appendix B
Federal Register / Vol. 45, No. 194 / Friday, October 3, 1980 / Rules and Regulations
ENVIRONMENTAL PROTECTION
AGENCY
40 CFR Part 125
[FRL 1609-1]
Ocean Discharge Criteria
AGENCY: Environmental Protection
Agency.
ACTION: Final rule.
SUMMARY: EPA is promulgating final
guidelines under section 403(c) of the
Clean Water Act. These guidelines will
be applied in issuing and revising
National Pollutant Discharge
Elimination System Permits for
discharges into the territorial seas, the
contiguous zone and the oceans.
DATES: These guidelines become
effective on November 3,1980.
FOR FURTHER INFORMATION CONTACT
Kenneth Farber, Office of Water
Regulations and Standards (WH-586),
Environmental Protection Agency, 401 M
Street, SW. Washington, D.C. 20460,
202-472-5746.
SUPPLEMENTARY INFORMATION:
I. Background
EPA is today promulgating revised
guidelines for determining the
. degradation of the territorial seas, the
contiguous zone and the oceans.
Pursuant to section 403(a) of the Clean
Water Act, no National Pollutant
Discharge Elimination System
("NPDES") permit for discharges into
these marine waters may be issued
when these guidelines are in effect
except in compliance with the
guidelines.
These guidelines are issued pursuant
to section 403(c)(l) which provides that:
The Administrator shall, within one
hundred and eighty days after enactment of
this Act (and from time to time thereafter),
promulgate guidelines for determining the
degradation of the waters of the territorial
seas, the contiguous zone, and the ocean,
which shall include:
(A) the effect of disposal of pollutants on
human health or welfare, including but not
limited to plankton, fish, shellfish, wildlife,
shorelines, and beaches.
B) the effect of disposal of pollutants on
• ..irine life, including the transfer,
concentration, and dispersal of pollutants or
their byproducts through biological, physical,
and chemical processes: changes in marine
ecosystem diversity, productivity, and
stability; and species and community
population changes;
(C) the effect of disposal of pollutants on
esthetic, recreation, and economic values:
(D) the persistence and permanence of the
effects of disposal of pollutants;
(E) the effect of the disposal at varying
rates, of particular volumes and
concentrations of pollutants;
(F) other possible locations and methods of
disposal or recycling -of pollutants including
land-based alternatives; and
(G) the effect on alternate uses of the
oceans, such as mineral exploitation and
scientific study.
On October 15,1973, EPA
promulgated combined regulations
implementing section 102{a) of the
Marine Protection. Research, and
Sanctuaries Act and section 403(c) of
the Clean Water Act. The primary focus
of these regulations was on the ocean
disposal of waste material, including
sewage sludges, liquid and solid
industrial wastes and dredged materials,
by dumping from moving vessels.
In practice, these regulations proved
unworkable in many respects as section
403 ocean discharge criteria. At the
same time, operating experience
demonstrated that the ocean dumping
regulations themselves required
revision. EPA therefore determined that
both the ocean dumping regulations and
the ocean discharged criteria should be
revised and published as separate
regulations. All reference to section
403(c) guidelines was deleted from the
revised ocean dumping regulations
which were promulgated on January 11,
1977 (42 FR 2468). However, the Agency
encountered substantial difficulty in
developing revised ocean discharge
guidelines, and, until recently, there
have been no published national
guidelines in place. Since withdrawal of
the original guidelines, permit writers
have been implementing section 403 on
a case-by-case basis.
On June 21,1979, the Pacific Legal
Foundation filed suit in United States
District Court for the Eastern District of
California, seeking, among other things,
that EPA promulgate revised section
403(c) guidelines, Pacific Legal
Foundation v. Costle, Civ. No. S-79-429-
PCW. The National Wildlife Federation
intervened in that lawsuit. On October1
31.1979, the Court ordered EPA both to
promulgate these guidelines and to
publish interim guidelines stating
Agency policy in reviewing, issuing, or
denying NPDES permits under section
403, pending promulgation of the final
guidelines. The interim guidelines were
published in the Federal Register on
November 15i 1979, 44 FR 65751, and
they will be superseded by the final
guidelines published today.
The Agency then published proposed
ocean discharge criteria in the Federal
Register (45 FR 9548) on February 12,
1980, held an oral hearing on the
proposal on March 21,1980, and
provided a comment period for
submission of written comments which
was to end on March 28,1980. At the
request of various interested groups, the
comment period was extended for 30
days. Based on the extended comment
period and on the large volume of
comments received, the Agency moved
the court to extend the final
promulgation date by 120 days beyond
the July 30 deadline. The court extended
the final deadline until September 30.
1980.
II. Development of the 403 Guidelines
1. Synopsis of the Guidelines
Section 403 is intended to prevent
unreasonable degradation of the marine
environment and to authorize imposition
of effluent limitations, including a
prohibition of discharge, if necessary, to
ensure this goal. These guidelines were
developed to satisfy this intent. They
provide flexibility to permit writers to
tailor application requirements, effluent
limitations, and reporting requirements
to-the specific circumstances of each
discharger's situation, while ensuring
consistency and certainty by imposing
minimum requirements, in situations
where the long-term impact of a
discharge is not fully understood.
Under these guidelines, no NPDES
permit may be issued which authorizes
a discharge of pollutants that will cause
unreasonable degradation of the marine
environment. Prior to permit issuance,
the director, defined as either the
Regional Administrator or the State
Director where there is an approved
State program, or an authorized
representative, is required to evaluate
whether a proposed discharge will cause
such degradation. In making this
determination, the director is to consider
the factors specified in § 125.122 (a) and
tb).
In cases where sufficient information
is available for the director to make a
reasonable determination whether
unreasonable degradation of the marine
environment will occur, the director is
governed by § 125.123 (a) and (b) of the
regulations. Discharges which will cause
unreasonable degradation will be
prohibited; other discharges may be
permitted under conditions necessary to
ensure that such degradation will not
occur.
In those cases where the director is
unable to determine whether
unreasonable degradation will occur,
§ 125.123(c) governs. No discharge in
this situation is allowed unless the
director can reasonably determine that:
(1) the discharge will not cause
irreparable harm to the marine
environment while further evaluation is
undertaken; (2) there are no reasonable
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alternatives to the discharge; and (3) the
discharge will comply with certain
mandatory permit conditions, including
a bioassay-based discharge limitation
and monitoring requirements. These
permit conditions will assist in
determining whether and to what extent
further limitations are necessary to
ensure that the discharge does not cause
unreasonable degradation. If further
limitations are necessary, § 125.123(d)(4]
provides that the permit must be then
modified to include these additional
limitations or else revoked.
These guidelines encourage the use of
available information in addition to any
supplied by the permit applicant. Thus,
the director may make determination
based on information such as that
contained in any relevant environmental
impact statement section 301(h) or other
variance applications, existing technical
and environmental field studies, or EPA
industrial and municipal waste surveys.
2, Relationship Between the Statute and
the Guidelines
(a) Section 403(c)(l}—Section 403(c)(l)
specifies seven factors which are to be
included in guidelines for determining
the degradation of marine waters. These
factors form the basis for the
determinations which must be made
pursuant to these guidelines.
Most of the statutory factors,
including 403(c)(l](A). (B), (C). (D), (E).
and a portion of (G), involve
consideration of the biological effects of
the discharge of pollutants. These
factors, either directly or indirectly,
must be evaluated by the director in
determining whether a discharge will
cause unreasonable degradation of the
marine environment Section 125.122(a)
requires that the director assess such
variables as the location of the
discharge, including the composition of
the biological community and existence
of special aquatic sites, such as marine
sanctuaries; the nature of the pollutants
which are to be discharged, including
their quantities, composition, potential
for bioaccumulation, persistence and
their transport in the environment and
the effect on human health. This
assessment should adequately address
the statutory factors relating to
biological effects of the discharge.
Section 403(c)(l)(C) also involves
consideration of economic and social
impacts of the discharge, as does section
403(c)(l)(G). The guidelines address
these factors in assessing whether a
discharge will cause unreasonable
degradation of the marine environment.
Section 122.121(f) defines "unreasonable
degradation of the marine environment"
to include, among other things, "loss of
esthetic, recreational or economic
values which are unreasonable in
relation to the benefit derived from the
discharge." Thus, even where the
director has determined that there are
no significant changes in ecosystem
diversity, productivity and stability, and
there is no threat to human health, he
may conclude that a discharge may not
be authorized if the adverse impact on
such activities as fishing, recreation,
and/or other economic or social benefits
is unreasonable in relation to benefits,
such as oil and gas production, derived
from the discharge.
Section 403(c)(l)(F) involves
consideration of other possible locations
and disposal methods for pollutants.
Although EPA has considered this factor
in developing these guidelines, the
director is not required to assess
alternatives in all cases. Under section
125.123(c)(2) the director must assess the
availability of alternatives, including
land-based alternatives, only in those
cases where he cannot determine that a
discharge will not cause unreasonable
degradation of the marine environment.
Additionally, the guidelines establish
a presumption that discharges in
compliance with sections 301(h), 316(a),
301 (g) and State water quality standards
under section 303 will not cause
unreasonable degradation. Although the
director may, on the basis of the factors
specified in § 125.122(a), conclude that
additional permit limitations are in fact
necessary even though the requirements
of these other sections have been met,
the similarity between the objectives
and requirements of these provisions
and those of section 403 warrants a
presumption that discharges in
compliance with these sections also
satisfy section 403. Also, even though
the regulations provide that a successful
section 316(a) demonstration creates
only a rebuttable presumption that
section 403 has been satisfied, the
provisions of section 316(c) may in some
cases preclude the imposition of more
stringent limitations under section 403.
POTWs obtaining section 301(h)
variances are entitled to a presumption
that their entire discharge is in
compliance with section 403. However,
the presumption applies only to the
thermal component of a discharge
subject to a 316(a) variance or to those
specific non-conventional pollutants
subject to a section 301(g) variance or to
pollutants specifically limited by criteria
in State water quality standards. Each
of those provisions, like section 403, is
geared toward assessing the
environmental impact of a discharge. In
order for a point source to receive a
section 301{g), 301(h] or 316(a) variance,
an evaluation of the biological and
environmental effect of the discharge is
required. Indeed, the statutory factors
specified in these sections are similar to
those contained in section 403(c).
Similarly, State water quality standards
established pursuant to section 303 of
the Act are designed to preserve the
quality of waters under State
jurisdiction, including the territorial
seas, and compliance with these
standards should insure protection of
the uses for which the waters are
designated with respect to pollutants for
which standards have been established.
(b)' Section 403(c)(2)—Section
403(c)(2) states that:
Where insufficient information exists on any
proposed discharge to make a reasonable
judgment on any of the guidelines established
pursuant to this section no permit shall be
issued under section 402 of this Act.
This section is the basis for two
central elements of these requlations.
First, the guidelines require that the
director make potentially complex
factual determinations on the basis of
information which, in many cases, may
be conflicting and in dispute. Section
403(c)(2) provides that the standard on
which the director is to make these
judgments is one of "reasonableness." In
assessing the information in the
administrative record, the director may
authorize the discharge of pollutants if
he is able to make a "reasonable
judgment" about the determinations
specified in the guidelines. Although
these issues may involve scientific
matters, the director is not bound by the
same burden of proof which a scientist
might require to reach a conclusion. The
administrative process and the burden
of proof in making these determinations
are discussed below.
Second, the regulation provides, as
required by section 403(c)(2), that the
• director may not authorize the discharge
of pollutants if there is insufficient
information to make these judgments.
The regulation does not, however,
require that there be complete
knowledge of the impact of a discharge
prior to permit issuance. Section
125.123(cJ provides that a permit may be
issued if the director has sufficient
information to reasonably conclude.
among other things, that the discharge
will not cause irreparable harm to the
environment while additional
information is collected. The provision
implements Congress' intent that
"discharges permitted today will not
irreversibly modify the oceans for future
uses." S. Rep. No. 92-414, 92nd Cong.,
1st Sess. at 75 (1971). It should insure
adequate protection of the environment
at all times while allowing the director
to issue NPDES permits where existing
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data may not be adequate to assess the
long term impact of a discharge.
3. The Role of Section 403(c) Guidelines
in NPDES Permit Issuance
These guidelines will be used to
develop NPDES permits for the •
discharge of pollutants into the
territorial seas, the contiguous zone and
the oceans. Application of the guidelines
will aid in protecting marine resources
and their uses from the impact of
pollution and in preventing
unreasonable degradation of the marine
environment.
Although sometimes described here as
"guidelines" or "criteria", these
promulgated regulations establish
minimum requirements on discharges to
protect the receiving waters. These
guidelines will be used in evaluating
applications for new, modified or
renewed permits as they are submitted.
These guidelines apply in addition to
other applicable provisions of the Clean
Water Act. Permittees subject to section
403 must still comply with all other
requirements of the Act'including
applicable technology-based
requirements specified by sections 301,
304 or 306 and water-quality based
limitations specified by sections 303 or
307. Permittees may in certain
circumstances be subject to the
provisions of section 311 as well.
Section 403 applies to all discharges
seaward of the inner boundary of the
territorial seas. This boundary is defined
by section 502(a) of the Act to be the—
belt of the seas measured from the line of
ordinary low water along that portion of the
coast which is in direct contact with the open
sea and the line marking the seaward limit of
inland waters. ...
This definition limits the number of
land-based dischargers subject to
section 403. For example, Chesapeake
Bay, Boston Harbor, New York Harbor,
San Francisco Bay and Puget Sound lie
inside this inner boundary so that
discharges into these waters are not
subject to section 403 requirements. Of
the approximately 62,400 existing
NPDES permittees, 232 are land-based
point sources discharging seaward of
this inner boundary. These include 102
publicly-owned treatment works, 74
industrial plants, 25 steam electric
plants and 31 federal facilities. These
figures do not include the dischargers in
Alaska whose location relative to the
baseline defining the ocean boundary is
not known.
In addition to these land-based
dischargers, section 403 applies to all
other point sources discharging into the
marine waters covered by this
regulation. By far the largest group of
marine dischargers are oil and gas
exploratory and production facilities.
The Agency estimates that there are
approximately 3,000 such facilities now
operating.
III. Modifications to the Proposal
EPA provided both an oral hearing
and a written comment period on the
proposal, the latter extended by thirty
days in response to the request of
several interested groups. The preamble
to the proposal specifically solicited
comment on certain points, including
mixing zone definition and
determination, control of toxic
pollutants; monitoring requirements and
procedures; and effect of meeting
requirements for a Section 301(h)
variance.
Ten persons testified at the March 21,
1980 hearing on the proposal, and
written comments were received from 81
parties, including many industrial
groups, municipalities, conservation
groups, federal agencies and several
State governments. A listing of these
commenters is presented in Appendix A.
The commenters addressed the issues
raised for comment in the preamble and
raised a range of additional issues
concerning EPA's approach in
developing the proposed regulations.
Detailed responses to the major public
comments are presented in Appendix B.
In conjunction with the public
comment review, EPA has reevaluated
the proposed regulations and concluded
that certain changes are appropriate.
The final regulations retain the basic
approach of the proposal. As in the
proposal, this regulation provides that
the director must determine whether a
discharge will cause unreasonable
degradation of the marine environment.
Based on review of the numerous
comments, however, EPA has made
certain modifications which are
intended to provide greater clarity to the
permit writer, to ensure consistency in
application of this regulation and to
minimize burdens on the permit
applicant.
Under the proposed regulation,
applicants were required to submit a
wide range of analyses and evaluations.
Numerous commenters objected, stating
that submission of this information was
unnecessary, and in many cases
redundant. In response to these
comments, the final regulations now
provide that the director may request
information from the applicant but is
encouraged to use other available
sources, such as environmental impact
statements, section 301(h) variance
applications, consolidated permit
applications, or EPA industrial and
municipal waste surveys.
These final regulations also clarify the
director's authority to issue permits
where certain pre-issuance
determinations relating to the effects of
a discharge cannot be made. These
regulations now provide that, in those
cases, the discharge of pollutants may
be authorized only where the director
has sufficient information to make
reasonable determinations regarding the
potential for irreparable harm from the
discharge and on the availability of
alternatives. These regulations also
establish certain minimum permit
requirements in these cases. These
requirements, identified as possible
permit conditions in the.proposed
regulation, have been made mandatory
in part in response to comments that the
regulations needed to provide greater
guidance to the director and applicant
regarding permit conditions.
Finally, with respect to the
relationship between the permit
requirements of section 403(c] and
variances issued under sections 301(g),
301(h), or 316(a), the final regulation
provides that an applicant who has met
the conditions necessary to receive such
a variance is presumed to be in
compliance with section 403(c) for those
pollutants to which the variance applies.
IV. Procedures for Issuance of Permits
Under Guidelines
1. Determination of Applicability of
Section 403
The threshold determination for
applicability of the section 403
guidelines is whether a proposed
discharge will occur seaward of the
inner boundary of the territorial seas.
EPA's consolidated permit regulations
(45 FR 33290, May 19,1980} require that
applicants list the latitude, longitude
and name of the receiving waters for
each outfall. Where the director is
uncertain as to whether the outfall is
within the waters covered by section
403(c), he should request guidance from
EPA headquarters. Where the proposed
discharge is in an area where the
baseline defining the boundary of the
territorial seas has not been determined,
EPA will request a determination from
the Department of State, which is
responsible for defining the boundaries •
of the territorial seas.
2. Determination of Information
Requirements Under Section 403(c)
Once a determination has been made
that section 403(c) applies to a particular
discharge, the director must determine
what information is required to evaluate
the discharge according to the section
403(c) criteria. The first thing that the
director should do is survey the
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currently available information about
the discharge itself and about the area
•jj which the discharge would occur.
This information would include'the data
contained in the consolidated
application form, as well as the data
available from Agency reports and
studies. The director should also notify
tjse applicant of the existence of other
currently available information.
After completing this survey of the
available information, the director
should determine what additional
information would be required from the
applicant for the evaluation of the
impact of the ocean discharge as
required by section 403(c). The applicant
will have the responsibility of collecting
this additional information and of
submitting it to the director.
3. Determination of Unreasonable
Degradation of the Marine Environment
Section 125.121(e) defines
"unreasonable degradation of the
marine environment" to include—
fl) significant adverse changes in
ecosystem diversity, productivity and
stability of the biological community within
the area of discharge and surrounding
biological communities; (2) threats to human
health through direct exposure to pollutants
or consumption of exposed aquatic
organisms; or (3) loss of esthetic, recreational,
identific or economic values which are
unreasonable in relation to the benefits
derived from the discharge.
Sections 125.122 (a) and (b) specify an
array of factors relevant in making these
determinations.
In many cases the director will be
able to reach conclusions based on data
related to the nature of the proposed
discharge. In areas which do not contain
sensitive species or unusual biological
communities or are not important for
surrounding biological communities, the
director may conclude that discharges
containing primarily conventional
pollutants will not cause unreasonable
degradation. This is especially
appropriate where the data indicate that
there will be significant mixing with the
receiving waters based on the flow of
we discharge and the physical
characteristics of the discharge site,
such as water depth and turbulence.
This determination may be appropriate
for such dischargers as small publicly
owned treatment works ("POTWs") and
for industrial dischargers such as fruit
canneries and fish processors.
For discharges into areas of biological
concern or for complex or toxic
discharges, additional evaluation may
°e necessary to determine whether a
Proposed discharge will cause
unreasonable degradation. In assessing
">e need for further evaluation, the
director should consider the
vulnerability of the area of discharge
and its role in the larger biological
community. Significant or sensitive
areas might include spawning sites,
nursery or forage areas, migratory
pathways or areas necessary for other
functions or critical stages in the life
cycle of organisms, areas of high
productivity, or areas under stress due
to biological or climatic conditions or
discharges from other sources.
Additionally, the director should
consider whether a discharge will affect
marine and wildlife species which are
identified as endangered or threatened
pursuant to the Endangered Species Act,
16 U.S.C. § 1531 et seq.r and those
species critical to the structure or
function of the ecosystem, such as in
food chain relationships.
An assessment of the potential
toxicity of a discharge should initially
focus on the pollutants which are
present in significant quantities relative
to marine water quality criteria
developed pursuant to section 304(a).
The potential for bioaccumulation or
persistence of pollutants in the
environment is of particular importance.
The director must also consider the
potential impacts of the discharge on
human health either directly as through
physical contact or indirectly through
the food chain. These factors should be
addressed when considering the
location of the discharge and the type
and volume of the discharger's effluent. •
Determinations of the economic
impact of the discharge should be based
on the potential effect of the discharge
•on such activities as commercial fishing,
recreation, mineral exploitation and
scientific study. In considering whether
a discharge will cause unreasonable
economic impacts, the social as well as
economic effects on a community should
be considered.
Much, if not all, of the information
necessary to make these evaluations
already will be available to the director.
Pursuant to section 122.53 of the
consolidated permit regulations,
applicants for NPDES permits must
submit a range of significant
information, including in many cases a
detailed analysis of toxic pollutants in
the waste stream. Additionally, any
relevant environmental impact
statements or section 301(h), 301(g) or
316(a) variance applications should
provide significant data about the
environmental impact of the proposed
discharge. EPA industrial and municipal
waste surveys and any data from
relevant technical and environmental
field studies that may have been
conducted may serve as a source of
information. Finally, Coastal Zone
Management Plans or proposals for
designation of an area as a marine
sanctuary should also contain relevant
information.
In cases where available information
is insufficient for the director to make a
determination, he may request
additional information of the applicant
pursuant to § 125.124. Where further
analysis of the area of the proposed
discharge is required, the director may
require the applicant to perform
assessments similar to those identified
in the Technical Support Document
prepared in conjunction with EPA's
section 301(h) regulations (44 FR 34784,
June 15.1979). EP'A headquarters will be
available to provide assistance to the
permit writer in developing information
requirements where the discharge may
be affecting an area of biological
concern.
The director ^should work with the
applicant to determine what types of
assessment are necessary and to review
or evaluate the assessment as it
progresses. This level of participation is
intended to determine the actual
feasibility and the costs of such
assessments for the applicant.
Furthermore, it should avoid duplicative
or inadequate assessments, thereby
preventing delays in permit issuance.
The Agency recognizes that some of the
anticipated assessments required for
permit issuance on the Outer
Continental Shelf are beyond those •
which can reasonably be expected of
the applicant and will require continued
Agency research efforts.
The guidelines establish a
presumption that discharges in
compliance with sections 301(g), 301(h),
316(a) or State water quality standards
will not cause unreasonable degradation
with respect to the pollutants covered
by those sections. Unless available data
indicate that a discharge will cause
unreasonable degradation, the director
need not take additional steps, including
the compilation of additional data, to
support a conclusion that no further
limitations on the discharge of these
pollutants is necessary.
4. Determination of Irreparable Harm
Section 125.123(c)(l) requires that the
director determine whether a discharge
will cause irreparable harm to the
marine environment in situations where
he cannot determine whether the
discharge will cause unreasonable
degradation. Although the concepts of
"irreparable harm" and "unreasonable
degradation" involve similar
considerations, the determination of
"irreparable harm" is much narrower in
scope.
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In assessing the probability of
"irreparable harm", the director need
not focus his analysis on the overall
impact of the discharge. Rather, he is
only required to make a reasonable
determination that the discharger,
operating pursuant to the permit
conditions established in § 125.123(c),
will not cause permanent and significant
harm to the environment during the
period in which further data on the
effects of the discharge are collected. In
cases where such data, primarily that
produced by monitoring, indicate that
continued discharge will produce
unreasonable degradation, the discharge
must be halted or additional limitations
established. Although evaluation of
irreparable harm may in some cases
involve difficult factual issues,
determinations of this kind are currently
a part of the NPDES permit issuance
process. Pursuant to 40 CFR
124.60(a)(2)(ii), the presiding officer at
an evidentiary hearing may authorize a
facility to commence discharging prior
to receipt of a final NPDES permit if the
permit applicant demonstrates, among
other things, that the discharge will not
cause "irreparable harm to the
environment . . ." This is essentially
the same finding which the director must
now make pursuant to these guidelines.
Certain factors are particularly
significant in assessing the likelihood of
"irreparable harm". Two such factors
are the quantity of pollutants expected
to be discharged and their potential for
persistence in the marine environment.
For example, a permit writer might
authorize the operation of exploratory
oil and gas facilities or a limited number
of production facilities based on a
conclusion that these limited operations
will not cause irreparable harm to an
area.
An additional factor is the sensitivity
of the area into which the discharge is
proposed. The director might conclude
that a discharge could cause irreparable
harm to unusual and interdependent
communities, such as the coral reefs and
associated communities of the Flower
Garden Banks proposed marine
sanctuary in the Gulf of Mexico. In such
areas special conditions, including the
prohibition of discharge, might be
required.
Finally, data on the effect of similar
discharges in similar areas is directly'
relevant to the determination of
irreparable harm. Information
demonstrating the recovery of the
environment after the cessation of
discharges from similar facilities would
be a strong indication that irreparable
harm is not likely to occur. For example,
data indicate that even in areas of
biological concern, biological
communities reestablish themselves
after the termination of discharges from
publicly-owned treatment works. Thus,
where the other provisions of
§ 12S.123(c) are satisfied, the director
might properly conclude that discharges
from POTWs pursuant to this section
may be authorized while further
information is being collected.
5. Determination of Reasonable
Alternatives
These guidelines establish two bases
for determining whether reasonable
alternatives to the proposed discharge
exist. The first is the physical
availability of alternative sites for
disposal of pollutants. Such alternative
sites might inciude either disposal
facilities located on land, discharge
point(sj within internal waters, or
existing ocean dumping sites approved
pursuant to the Marine Preservation,
Research and Sanctuaries Act. In
determining whether a site is a
reasonable alternative to on-site
disposal, the director should consider its
distance from the site of the proposed
discharge and whether its use would
cause unwarranted economic impact on
the discharger. For example, shipping
wastes over long distances would likely
impose such impact. This provision is
intended to ensure some rule of reason
in evaluating alternatives, it is not
intended to impose a "cost/benefit"
analysis of alternative sites.
In considering the availability of
alternatives the director shall consider,
based on available information or that
requested from the applicant, the
estimates of the amount of material
requiring disposal. He should review the
availability of existing land-based
disposal sites and ocean dumping sites
within a reasonable distance from the
point of discharge and the estimated
uncommitted capacity of such sites. The
director should evaluate any reports of
economic impact of discharge
alternatives as may be supplied by the
applicant.
The second basis for evaluating the
feasibility of alternative sites relates to
the relative environmental harm of
disposal. Pursuant to section 121(e)(2),
alternative disposal sites are not
considered "reasonable alternatives" if
on-site disposal is judged to be
environmentally preferable. Thus, the
discharge of pollutants might be
authorized where disposal in alternative
sites might produce equal or greater
environmental harm than on-site
discharge, or where transportation to
alternative sites might produce a
significant risk of greater environmental
harm or a significant risk to human
safety. For example, during certain
seasons it may be undesirable to
transport wastes off-site in areas of the
•North Atlantic or Alaska. Where the
environmental or human health risks of
transportation are significant, such
transportation should not be considered
a reasonable alternative.
6. Determination of Permit Conditions
Section 125.123(d) identifies specific
permit conditions which are required for
the issuance of a permit where a pre-
permit issuance determination regarding
degradation of the marine environment
cannot be made. The director may also
require any necessary permit conditions
identified in section 125.123(d) to assure
that unreasonable degradation of the
marine environment will not occur
under § 125.123(a).
(a) Limiting Permissible
Concentration Requirements—Section
125.123(d)(l) requires, if a determination
regarding unreasonable degradation
cannot be made, that the discharge must
pass certain bioassay-based
requirements similar to those of EPA's
ocean dumping regulations (40 CFR Part
227).
The applicant must demonstrate that
his discharge will not exceed the
limiting permissible concentration
("LPC") at the boundary of the mixing
zone for a liquid phase and a suspended
particular phase bioassay, in
accordance with procedures for
determining the LPC which are
described in Bioassay Procedures for
the Ocean Disposal Permit Program,
U.S. EPA 600/9-78-010 March 1978 and
in Ecological Evaluation of Proposed
Discharge of Dredge Material into the
Ocean Waters, EPA/Corps of Engineers,
July 1977. If these manuals are revised in
the future, bioassays shall be performed
in accordance with any such revisions.
These regulations require an LPC
which is derived from, but not identical
to, the ocean dumping bioassay
requirements. First, these regulations do
not use section 304(a)(l) marine water
quality criteria as a basis for
determining an LPC. By use of a
bioassay-based LPC, the ocean
discharge criteria address the impact of
the whole effluent and account for any
synergistic or antagonistic effects. EPA
recognizes that section 304(a)(l) criteria
may in some cases require changes to
reflect site-specific conditions, and the
Agency is Devaluating the use of marine
water quality criteria in the ocean
dumping program. , .,
The ocean discharge criteria also use
a mixing zone extending laterally 100
meters in all directions from the
discharge point(s) or to the boundary of
the zone of initial dilution as calculated
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65947
hy a plume model approved by the
director, whichever is greater, unless the
jjrector determines that the more
restrictive mixing zone or another
definition of the mixing zone is more
appropriate. In calculating the dilution
it the boundary of the mixing zone, the
discharger may use any of the various
documented plume models and
dispersion models appropriate for the
discharge and approved by the director.
Some of these models are referenced in
the technical documents for EPA's ocean
dumping regulations and the technical
document for EPA's section 301(h)
regulations.
Where the discharge contains a solid
phase, the applicant will be required to
perform the solid phase bioassay and
bioaccumulation testing on.the waste
material in accordance with procedures
described in Ecological Evaluation of
Proposed Discharge of Dredge Material
into the Ocean Waters. EPA/Corps of
Engineers, July 1977. For example, if a
bioassay analysis is required in the case
of offshore oil and gas platforms, the
solid phase bioassay would be
conducted as a test on drilling muds and
cuttings which are to be discharged.
Not all applicants may have to
perform bioassay tests on their effluents.
Applicants may submit bioassay
analyses performed on other wastes if
the applicant provides documentation to
«how that the composition of the waste
analyzed typifies that which the
applicant is discharging or intends to
discharge.
(b) Monitoring Requirements—Where
a pre-issuance determination regarding
degradation of the marine environment
cannot be made, § 125.123(d)(2) requires
that a monitoring program be in place
which is sufficient to assess the impact
of the discharge on water, sediment, and
biological quality including, where
appropriate, analysis of the
bioaccumulative and/or persistent
impact on aquatic life of the discharge.
This monitoring program may include
effluent analysis, bioassay analysis and
field studies. The technical document Tor
EPA's section 301(h) regulations should
provide support in developing such a
monitoring program. It is not possible to
oake an a priori determination as to
what constitutes an acceptable
monitoring program. Site-specific
conditions such as the size of the
discharger's waste stream, the types of
Pollutants discharged, and the location
of the discharge will play a role in
determining what if any specific
Monitoring will be required under
section 403(c) in addition to other
"PDES monitoring requirements.
Section 125.123(d)(2) provides the
director a flexible mechanism to develop
such site-specific monitoring
requirements. For example, a low
volume discharger whose waste stream
is unlikely to contain significant
amounts of toxic pollutants will not be
required in most cases to establish a
monitoring program under these
regulations. Similarly, a discharger of
pollutants into an area of biological
concern may be subject to more
stringent monitoring requirements than
one not discharging into such an area.
Monitoring programs, in some
instances, may be coordinated for
several dischargers. For example, with
offshore oil and gas platforms, areawide
monitoring programs for several
dischargers may be the desirable
monitoring approach. EPA headquarters
has been active in assisting the regions
in developing monitoring programs for
offshore oil and gas exploration in areas
of biological concern such as the Flower
Garden Banks and Georges Bank. Those
monitoring programs will serve as
valuable guides for the development of
additional monitoring programs for
other areas of offshore oil and gas
exploration and production. EPA
headquarters will continue to play an
active role in providing technical
assistance in developing such
monitoring programs.
(c) Other Permit Conditions—Under
§ 125.123(d)(3), the director may also
require under other permit conditions on
the discharge. For example, the director
may require seasonal restrictions on the
volume of wastes discharged where
such restrictions are needed to assure
protection of the marine environment.
Seasonal restrictions may be necessary
where the discharge is itself affected by
seasonal conditions or where the
biological community may become more
sensitive to the impact of the discharge
during certain seasonal conditions, such
as during migration or spawning.
The director may require that the
applicant perform bioaccumulation
testing of the liquid and/or suspended
particulate phase of the discharge where
the director suspects such potential for
bioaccumulation may exist, based oiv
the nature of the pollutants discharged.
The director may also require process
modifications, such as the substitution
of less hazardous chemicals for those
which are potentially harmful. He may
also require process changes which
would favor the recycling and reuse of
potentially harmful pollutants. The
Agency has recently established a task
force to evaluate the discharges from
offshore oil and gas exploration and
production facilities and to evaluate
alternate control strategies to mitigate
the effects of such discharges, which
include drilling muds and cuttings- and
produced water. Its recommendations
may be used in drafting future
requirements under section 403(c)
authority.
The director may also require that
diffuser systems for the discharger be
sufficient to assure adequate dispersion
of the waste stream.
7. The Administrative Process and
Burden of Proof
Under the Act and this regulation, the
director is responsible for making
"reasonable judgments" on the
preceding issues, and these judgments
will be made oil available information
compiled in the administrative record of
the permit issuance. As discussed
above, this information may come from
many sources including data submitted
under the consolidated permit
application form, environmental impact
statements or section 301(h) variance
applications. These guidelines do not
require that all applicants submit
specific information to support the
section 403 determinations, and the
director is encouraged to make use of
existing information not prepared by
applicants.
However, under the Clean Water Act,
the Administrative Procedure Act, and
EPA's consolidated permit regulations it
is the applicant who is responsible for
persuading the Agency that a permit
should be issued. See 40 CFR
124.85(a)(l) and Opinion of the General
Counsel No. 72. This obligation is
particularly apparent with respect to
applicants seeking permits to discharge
into marine waters. Section 403(c)(2)
requires that the director deny an
NPDES permit application if there is
insufficient information to make
reasonable judgments under the
guidelines. This means that the permit
applicants should be prepared to submit
sufficient information to support a
determination to issue an NPDES permit.
Under the Agency's permit issuance
procedures there is opportunity to
submit information for the
administrative record. An applicant or
interested person who disputes any
permit condition or tentative decision to
deny an application must submit
available information supporting their
position during the public comment
period. 40 CFR 124.13. In any subsequent
evidentiary hearing on the permit, the
Agency will have the burden of going
forward to present its case supporting a
challenged permit condition, but, at the
conclusion of the Agency's presentation,
the applicant or any other hearing
participant has the burden of going ••
forward to present its case. 40 CFR
124.85(a) (2) and (3). Moreover, the
ultimate burden of persuading the
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Agency to issue a permit remains at all
times on the permit applicant.
V. Cost and Economic Impacts
Executive Order 12044, 43 FR 12661
(March 23,1978), requires EPA and other
agencies to perform Regulatory
Analyses of certain regulations. EPA's
plan for implementing Executive Order
12044, 44 FR 30988 (May 29,1979),
requires a Regulatory Analysis for major
regulations involving annual compliance
costs of S100 million or meeting other
specified criteria. Where these criteria
are met, EPA's implementation plan
requires a formal Regulatory Analysis
including an economic impact analysis
and an evaluation of regulatory
alternatives. The Agency has
determined that none of the criteria for
requiring a regulatory analysis has been
exceeded and therefore, the
promulgated regulations for ocean
dischargers do not require a formal
Regulatory Analysis. Nevertheless, EPA
performed an analysis that does meet all
the requirements of Executive Order
12044 and EPA's plan for its
implementation.
In accordance with the requirements
under section 2(d)(8) of Executive Order
12044, the Agency intends to review the
effectiveness and need for continuation
of the provisions contained in this action
no more than five (5) years from the
effective date of these regulations. In
particular, we will solicit comments
from affected parties with regard to
actual costs incurred and other burdens
associated with compliance and will
also review environmental data to
evaluate the effectiveness of the
regulation after it has gone into effect.
EPA's economic analysis divides the
affected dischargers under the proposed
regulation into five subcategories:
POTWs, industrial dischargers, electric
utilities, federal facilities, and offshore
oil and gas wells. This analysis has
assessed unit price increases,
production changes for industrial
dischargers, and user cost increases at
POTWs.
The total cost of compliance is
expected to be $13 million in 1981 and to
increase to $28 million by 1985, in 1980
dollars. Costs may increase in-
succeeding years. The Agency expects
no significant economic impacts will
result from this regulation.
1. POTWs.
There are presently 102 POTWs
discharging 2.1 billion gallons of effluent
a day into the ocean. These POTWs are
currently operating under EPA's NPDES
regulations. Under these regulations
POTWs may be required to monitor,
perform benthic analyses.
bioaccumulation tests and run further
analyses of disposal alternatives in
addition to those required under their
present NPDES permit. The Agency
performed an economic analysis to
determine the potential costs and
increases in user charges currently paid
by households serviced by affected
POTWs. EPA estimates that 47 POTWs
will incur additional costs, due to their
location and the size of their discharges,
consisting of a first year cost of S1.2
million and an average annual cost of
$.828 million thereafter.
For 46 of the 47 affected POTWs. user
charges will increase between S.09 to
S.83 per family per year. The impact
analysis then compared these costs to
median family incomes and found that
no significant economic impacts would
occur. However, the impact analysis has
indicated that for one community user
costs could increase by S33.00 per family
per year.
Currently 36 POTWs subject to these
regulations have applied for 301(h)
variances. EPA has not yet begun to
issue decisions on section 301(h)
variance requests. However, much of the
information generated for purposes of
section 301(h) applications can be
utilized in determining compliance, with
the requirements of these regulations.
Furthermore, in recent years there has
been a trend towards centralization of
POTWs in many coastal areas. This
continued centralization will reduce the
number of ocean outfalls, thus lowering
total monitoring and user costs.
2. Industrial Dischargers
Industrial dischargers will face the
same type of compliance requirements
as POTWs. Monitoring and compliance
requirements for industrial dischargers
are dependent on the particular
geographic area as well as the
composition and volume of the effluent.
At the present time there are 74
industrial operations affected by this
regulation, discharging approximately
212 million gallons per day of effluent.
The Agency expects that small
industrial plants discharging non-toxic
pollutants will not be affected by this
regulation. EPA estimates 46 dischargers
will incur additional costs due to this
regulation, with a first year cost of S4.72
million, and annual costs of $3.52 million
in the following years.
An analysis was conducted for a
sample of industrial dischargers on both
the East and West Coasts to determine
the potential price increases that could
result due to this regulation. EPA
estimates that average unit prices for
products produced will generally
increase less than .1 percent to comply
with this regulation. No plant closures,
unemployment or other significant
economic impacts are expected due to
these requirements.
3. Federal Facilities
At the present time there are 31
federal facilities affected by this
regulation. These facilities are
discharging approximately 124 million
gallons per day of effluent into the
ocean. The greater part of the total, 75
million gallons per day, originates from
the strategic oil reserve construction site
on the Texas coast. The remaining 49
million gallons a day are from a variety
of small sources, e.g., Defense
Department and Coast Guard
installations. The Agency estimates that
four federal facilities discharging 102
million gallons per day will actually
incure additional costs from this
regulation. The total annual costs for
compliance under the proposed
regulation is expected to be
approximately S.476 million, with tne
largest proportion of this amount being
related to the construction of the United
States strategic oil reserve. EPA does
not expect any significant economic
impacts to occur due to expenditures by
these facilities.
4. Electric Utilities
The Agency does not expect any
significant costs to be incurred by
electric utilities. Compliance with the
present effluent limitation requirements
and with regulations implementing
section 316(a) of the Clean Water Act
are expected to result in compliance
with requirements in this regulation. The
Agency expects that monitoring for
chlorine discharges may be required at
some facilities. However, the cost of
such monitoring would not be
significant, and no economic impacts are
expected to occur.
5. Offshore Oil and Gas Operations
There are presently fewer than 30 oil
and gas platforms which are expected to
incur additional costs due to this
regulation. The Agency estimates that
7,582 exploratory and production wells
will be drilled between 1981-1985 with
approximately 835 (11 percent) expected
to incur additional costs resulting from
compliance with this regulation. The
Agency has based its assessment on the
assumption that compliance with
applicable NPDES permit requirements
will generally result in compliance with
these regulations for all oil and gas
wells except those located in areas of
.biological concern. Wells that cannot
meet the requirements of this regulation
through compliance with their NPDES
permit terms will be required to initiate
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65949
monitoring, testing, or changes in their
discharge practices.
The Agency's evaluation of the
economic effects of this regulation
assumed the installation of "zero
discharge" technologies in order to
evaluate the maximum possible impact
of these regulations. The cost of "zero
discharge" varies according to
geographic location, differences in
weather conditions, water depths,
biological communities, and other
similar factors. The economic analysis
groups wells into four regions—the
Atlantic Ocean, Gulf Ocean, West
Coast, and Alaskan waters.
The number of wells that would need
to make expenditures beyond those
required under existing NPDES permit
requirements was estimated from
Department of Interior data regarding
current and future lease tracts in the
ocean and from current trends in new
drilling. The estimates here project
activity from 1981-1985. Should the
amount of new drilling increase or
decrease beyond that time the annual
cost of this regulation would increase or
decrease proportionately.
It is estimated that the annual cost for
offshore oil and gas operations locating
in or near areas of biological concern
will incur compliance costs ranging from
seven to 23 million dollars per year
between 1981 and 1985. Costs for
various forms of monitoring are
expected to be less. There are numerous
alternatives which include process
changes, mud substitution and shunting
as potential compliance alternatives. A
number of combinations are possible,
depending on geographic location, water
depth, temperature and specific
biological life. For this reason only the
worst case, "zero discharge"
requirement is presented here.
The typical compliance cost per year
for each geographical area for the no
discharge alternative is presented
below. Costs for operations in the
Atlantic Ocean are expected to range
between .6 and 6.9 million dollars per
year, and Gulf of Mexico operations will
face costs ranging between 1 and 3.6
million dollars per year. Operations
located on the West Coast will face
costs ranging between 4.3 and 12.4
million dollars per year. However, the
California Ocean Plan requires outer
continental shelf operations to conform
to strict State requirements, which may
reduce the incremental compliance costs
under this regulation. Alaskan
operations will face compliance costs
ranging between 1.2 and 4.5 million
dollars per year.
Oil and gas prices at the well head are
not expected to be affected by this
regulation, since compliance costs
cannot be directly passed forward due
to various price controls. However, the
cost of this regulation may be
manifested in reduced bids for new
lease tracts. The net effect of this
regulation would be a loss in future
revenues to the federal government in
the amount which this regulation costs
the oil industry.
Dated: September 26.1980.
Douglas M. Costle,
Administrator.
Appendix A—Public Comments
The following parties responded with
comments regarding the February 12,
1980 Ocean Discharge Criteria
postmarked on or before the April 28,
1980 close of the public comment period:
Charles A. Lunsford, Commonwealth of
Virginia, State Water Control Board;
State of Hawaii, Dept. of Health; County
of San Diego, Community Services
Agency, Dept. of Sanitation & Flood
Control: City of Los Angeles, California
Dept. of Public Works: Menasha
Corporation; National Manufacturing
Company; Commonwealth of Virginia,
State Water Control; County Sanitation
Districts of Los Angeles County; Crown ,
Zellerbach Environmental Services;
Kaiser Aluminum & Chemical
Corporation; Boise Cascade, Paper
Group; Davies Hamakua Sugar
Company; Dept. of Health, Education &
Welfare, Public Health Service;
Hawaiian Sugar Planters' Association;
Hilo Coast Processing Company;
Netarts-Oceanside Sanitary District;
International Paper Company; State of
Alaska, Dept. of Fish & Game;
Commonwealth of Virginia, Hampton
Roads Sanitation District; Marathon Oil
Company, Production Operations; San
Francisco Wastewater Program, City
and County of San Francisco, California;
U.S. Cape May County Municipal
Utilities Authority, New Jersey; Dept. of
the Army, South Atlantic Division,
Corps of Engineers; State of California,
Resources Agency, Dept. of Fish and
Game; National Fisheries Institute, Inc.,
Natural Resources Defense Council, Inc.,
Sussex County Council, Georgetown,
Delaware; American Paper Institute/
National Forest Products Association,
Environmental Program, Houston
Audubon Society; Texas Eastern
Transmission Corporation; State of
Delaware, Department of Natural
Resources and Environmental Control,
Division of Environmental Control;
Department of the Air Force,
Engineering and Serivce Center; Star-
Kist Foods, Inc.; State of California,
Resources Agency, State Water
Resources Control Board; Office of the
Assistant Secretary of Defense, Energy,
Environment and Safety; the Ocean
County Utilities Authority, New Jersey,
National Wildlife Federation;
Department of the Army, Office of the
Chief of Engineers; Shell Oil Company;
Texaco, Inc., American Cyanamid
Company; Atlantic Richfield Company;
E. I. du Pont de Nemours and Company,
Inc.; Public Service Company of New
Hampshire; Chevron U.S.A., Inc.,
Environmental Affairs; Commonwealth
of Puerto Rico, Puerto Rico Aqueduct
and Sewer Authority; Exxon Company,
U.S.A.; Offshore Operators Committee,
Southern California Edison Company;
Virgin Islands Rum Industries, Ltd.,
Fried, Frank, Harris, Shriver &
Kampelman; Chevron U.S.A. Inc.,
Pillsbury, Madison & Sutro; City of
Watsonville, California; Conoco, Inc.;
Conservation Law Foundation of New
England, Inc.; Mobil Oil; Corporation;
Western Oil & Gas Association; Alaska
Lumber and Plup Company, Inc.,
Robertson, Monagle, Eastaugh &
Bradley; American Petroleum Institute;
Cody Biggs; Chemical Manufacturers
Association, Covington & Burling; City
of Skagway, Alaska, Robertson,
Monagle, Eastaugh & Bradley; Columbia
Gas System Service Corporation;
Department of Energy; Gulf Oil
Exploration & Production Company;
National Food Processors Association;
Pacific Legal Foundation; Phillips
Petroleum Company; Tuna Research
Foundation, Inc.; Union Oil Company of
California; U.S. Department of
Commerce, National Oceanic and
Atmospheric Administration,
Environmental Research Laboratories;
Natural Resources Defense Council, Inc.;
U.S. Fish & Wildlife Service, Alaska
Area Office; Utility Water Act Group,
Hunton, & Williams; Department of
Water & Power, the City of Los Angeles,
California; U.S. Department of the
Interior, Geological Survey.
The following parties responded with
comments postmarked after the April 28,
1980 close of public comment period:
U.S. Department of the Interior;
Chevron, U.S.A., Pillsbury, Madison &
Sutro; University of Southern Maine for
State Planning Office, State of Maine;
Monmouth County Board of Health,
New Jersey; State of Maine, State
Planning Office; Western Oil and Gas
Association.
The following parties testified at the
March 21,1980 hearing: George P. Haley,
Chevron USA; Frank Parker,
Coordinator, Environmental and
Government Affairs Chevron USA;
Elizabeth F. Kroop, Counsel for the
National Wildlife Federation; Walter J.
Zizik, Project Coordinator, South
Monmouth Regional Sewerage
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Appendix B
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Authority; Curt D. Rose. Manager,
Aquatic Sciences Division, Energy
Resources Company: Frank Melone,
Southern California Edison Company;
William A. Anderson. Attorney, Utility
Water Act Group; Joseph F. Dietz,
Coordinator of Environmental Affairs,
San Diego Gas and Electric Company;
Edward G. Gladbach, Civil Engineer,
Department of Water and Power, City of
Los Angeles: Peter Holmes, Research
Assistant. Atlantic Coast Project,
Natural Resources Defense Council.
Appendix B—Response to Public
Comments
1. Comment. The Agency received
several public comments questioning the
accuracy of the inventory in the
proposal and noting that there was
uncertainty over the exact location of
the baseline marking the landward
boundary of the territorial seas,
particularly in parts of Alaska, Florida,
Puerto Rico, Oregon and Washington.
Response. Even where the baseline
has not been plotted, there are available
nautical charts for the various bays and
harbors in question. The Office of the
Geographer in the Department of state is
responsible for the charts plotting
closing lines across islands and
shoreline markers depicting the baseline
of the territorial sea. That office is
assisted by the Interagency Baseline
Committee, chaired by the Department
of State, with other members coming
from the National Oceanic and
Atmospheric Administration, the Coast
Guard, and the Departments of Justice
and the Interior. The Committee meets
several times a year to make baseline
determinations.
To facilitate the ongoing
implementation of section 403, EPA has
identified dischargers whose coverage
under that provision is in question, due
to uncertainty concerning the location of
the baseline. The Agency has submitted
a written rquest to the Department of
State and the Interagency Baseline
Committee for a determination whether
these dischargers are outside the
baseline of the territorial seas and thus
subject to section 403. EPA will continue
to seek determinations when NPDES
permits are issued, modified, or
reissued, where there is doubt as to
whether a discharger is within the
purview of section 403.
2. Comment. One commenter stated
that while it appeared that the Agency
intended for coastal electric utilities to
be subject to the regulations, these
plants were not counted among the 71
land-based industrial dischargers.
Response! Electric utilities outside the
baseline are covered by this regulation
and EPA. in response to the comment.
confirmed this with representatives of
the affected industry during the public
comment period. The Agency also
extended the comment period by thirty
days, at the request of this commenter
and others, to allow additional time for
submission of comments on the
proposal.
The Agency has identified 25 covered
plants, which are included in the cost
and economic impact analysis for the
regulation. In addition, EPA has updated
the inventory of subject marine
dischargers. As the preamble notes, the
number of land-based dischargers
subject to section 403 is limited. The
updated inventory identifies 232 such
dischargers..including 102 POTWs, 74
industrial facilities, 25 steam electric
utilities, and 31 federal facilities. These
figures do not include dischargers in
Alaska whose location relative to the
baseline defining the boundary of the
territorial seas has not been established.
However, the Agency believes that most
of these dischargers are small and that
any environmental or economic impacts
would be minimal. The Agency also
estimates that there are some 3,000
subject offshore oil and gas platforms.
3. Comment. A number of commenters
stated that under the proposed
regulation, ocean dischargers might be
subject to more stringent controls and,
accordingly, might incur higher costs
than would dischargers into potentially
more sensitive estuarine and freshwater
systems where the assimilative capacity
of the body of water may be less than in
the oceans and the potential impact of
pollution on the ecosystem greater.
Response. As the preamble to the
regulation notes, the Clean Water Act
limits the coverage of section 403 to
dischargers into waters seaward of the
baseline marking the territorial seas.
This additional assurance of protection
and its limitation to the waters of the
territorial seas, the contiguous zone, and
the oceans is, therefore, a matter of
Congressional mandate. EPA has
designed its regulations to provide this
protection, as Congress has directed. As
to the question of costs, the Agency
anticipates that in most cases,
technology-based effluent limitations
required under other provisions of the
Act will be adequate in themselves to
afford the necessary protection for the
marine environment. As to freshwater
and estuarine systems, the Agency
agrees that these waters must be
protected also; the statutory authority to
accomplish this, however, rests in other
sections of the Act and in other
environmental statutes.
4. Comment. A number of commenters
suggested that the ocean discharge
criteria should merely have the effect of
guidelines, rather than regulatory
requirements, and should provide
flexibility and allow for discretion on
the part of the director to apply
appropriate portions of the guidelines to
the situation of an individual discharger.
Response. As noted in the preamble,
these regulations, although from time to
time described as "guidelines" or
"criteria" to avoid repetition, establish
mandatory requirements authorized by
section 403(c). Whatever the
terminology, they have the effect of
mandatory regulations because, at any
time that promulgated guidelines are in
effect, no NPDES permit may be issued
"except in compliance with such
guidelines." Nevertheless, the regulation
provides the permit writer flexibility in
tailoring information requests and
permit conditions to the circumstances
of individual dischargers, based on local
conditions.
5. Comment. Several commenters
expressed the concern that while the
proposed regulation required a permit
applicant to demonstrate to the director
that its discharge would have no
unreasonable adverse impact on the
environment, or that adequate toxics
control and monitoring programs were
in place, the proposal failed to tell the
applicant how to make such a
demonstration.
Response. In order to ensure "that
applicants will receive adequate
guidance, the final regulation has been
clarified to require that the director
inform an applicant of any specific
information that must be supplied. In
addition, in an attempt to minimize the
information collection obligation of
applicants, the final regulation provides
that the director may consider
information already available to him in
making the determinations required
under § 125.123(a), (b), or (c).
6. Comment. Several commenters
suggested that, for the territorial seas,
the purposes of section 403 were being
served already by State water quality
standards required under section 303
and by technology-based effluent
limitations under sections 301 and 304 of
the Clean Water Act. Another
commenter stated that pretreatment
programs required under section 307
should also ensure protection of the
marine environment.
Response. The Agency agrees that
State water quality standards,
technology-based limitations and
pretreatment programs are all necessary
to protect the marine environment.
While in most instances discharges in
compliance with such standards, limits,
and programs will also be determined to
pass the "no unreasonable degradation
test in these regulations, there may be
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65951
^stances where this will not be the
case. For example, there may be
instances where no State water quality
standards have been established for
specific pollutants being discharged.
Further. State water quality standards
do not generally apply beyond the limits
of the territorial seas, while the section
403 criteria apply also to the contiguous
zone and the oceans. In addition, there
nay be instances in which technology-
based controls will not be sufficient to
assure protection of a particular marine
environment, necessitating more
stringent controls to assure that the
section 403 criteria are met. Such may
also be true of pretreatment programs;
while the director may consider the
effectiveness of a given pretreatment
program in making the "unreasonable
degradation" determination, he should
not assume that the existence of
pretreatment ensures protection of the
marine environment for purposes of
section 403.
7. Comment. Several commenters
stated that while the proposed
regulation differentiated special "areas
of biological sensitivity" to assure those
areas were not adversely affected, the
regulation failed to adequately define
what constitutes such areas. Another
commenter suggested that the term
should be replaced by "areas of
biological concern" because "biological
sensitivity" connotes a narrow concern
for unique or fragile ecosystems. The
commenter suggested that the emphasis
of the regulation should be on
unwarranted ecological damage
regardless of the biological sensitivity of
the area.
Response. The scope of this regulation
is broader than protecting only those
areas that are termed "sensitive"; as in
the proposed regulation, these
guidelines seek to prevent unreasonable
degradation of the marine environment
regardless of where the discharge
occurs. Although the regulation no
longer attempts to classify areas as
"sensitive" or "nonsensitive", the
location of the discharge is an important
element in determining the level of
control necessary to prevent such
degradation. Section 125.122 identifies
for the director a number of factors
relating to the biology of the local
community which are important in
assessing the impact of a discharge.
8. Comment. A substantial number of
comments submitted on behalf of
various dischargers suggested that the
dischargers in question—including small
POTWs, electric utilities, seafood
processors, Alaskan logging operations,
and offshore oil and gas exploration and
Production wells—should be exempt
from the requirements of section 403.
Several commenters made the argument,
in some cases based on submissions of
technical data and reports, that their
discharges already were subject to
controls adequate to protect the marine
environment. Some stated that their
discharges resulted in only de minimus
effects on the environment. Some stated
that compliance with various provisions
of the proposal would result in economic
hardship.
Response. EPA has concluded that
there is no basis for categorically
exempting classes of subject dischargers
from the coverage of section 403. While
the data submitted by some commenters
may be useful in determining whether a
particular discharge will meet the
"unreasonable degradation" test, it does
not provide a basis for such a blanket
exemption. However, while a permit
writer is not precluded from seeking
additional site-specific information, the
submission of large quantities of data
for particular dischargers or classes of
dischargers makes it unlikely that a
permit writer will find it necessary to
require these applicants to submit any
substantial quantity of additional data.
Similarly, in the cases of small POTWs
and others where the discharge is
expected to have only a minimal impact,
the flexibility which the final regulation
provides will allow the permit writer to
take this situation into.account, rather
than mandating a rigid across-the-board
application of all requirements, with
their associated costs.
9. Comment. Several commenters
suggested that POTWs granted section
301(h) variances from secondary
treatment requirements should be
exempt from section 403 because of
significant similarities in the two
provisions. Another commenter,
however, stated that section 301(h)
contains no analogue to section
403(c)(l)(F) or (G) and asserted that
toxic pollutants are not adequately
controlled under section 301(h).
Response. Despite differences in
statutory language, sections 403 and
301(h) share similar objectives in
seeking to assure protection of the
marine environment, and the respective
determinations whether those objectives
have been met under each provision is
based on similar information. Section
301(h)(2) requires that a successful
applicant for a variance demonstrate,
among other things, that "such modified
requirements will not interfere with the
attainment or maintenance of that water
quality which assure protection of
public water supplies and the protection
of shellfish, fish, and wildlife, and
allows recreational activities in and on
the water." Section 125.61 of EPA's
section 301(h) regulations requires full
and detailed descriptions of the physical
characteristics of the discharge, its
biological impact on the marine
environment, and its impact on public
water supplies and recreation. Given
therefore that a successful section 301(h)
applicant will have collected and
presented substantial amounts of data
on the effect of its discharge on the
marine environment, including its
inhabitants and uses, the final ocean
discharge regulations provide that a
successful section 301(h) demonstration
creates a rebuttable presumption that an
applicant will satisfy the section 403(c)
guidelines as well. While a permit writer
is not precluded from placing additional
requirements on such an applicant under
these regulations, it is unlikely that this
will be necessary in light of the through-
going demonstration the applicant has
made for purposes of section 301(h).
This approach is consistent with
legislative history to the effect that
section 301 (h) applicants must comply
also with section 403. This language
indicates that Congress did not intend
for section 403 to become a dead letter
with the subsequent enactment of
section 301(h). Unlike the approach of
those commenters who sought to make
compliance with section 403 automatic
for an applicant who had obtained a
section 301(h) variance, the "rebuttable
presumption" approach does not treat
section 403 as redundant. Nor, however,
does it impose a redundant data-
gathering task on successful section
301(h) applicants either, taking account
as it does of the unmistakable
similarities in the showings required
under the two provisions.
The Agency disagrees with the
comment that the toxic control
provisions of the section 301(h)
regulations are not adequate. Moreover,
if a permit writer determines that toxic
pollutants in the discharge of a
successful section 301(h) applicant are
not adequately controlled for purposes
of section 403, he can require additional
controls or, if necessary, require zero-
discharge permit terms for those
pollutants.
10. Comment. Several commenters
suggested that coastal steam electric
utility plants granted a variance under
section 316(a) of the Act should be
exempt from demonstrating compliance
with section 403, on the grounds that the
demonstration necessary for obtaining a
section 316(a) variance provides the
requisite assurance that the marine
environment is protected for purposes of
section 403(c).
Response. To obtain a section 316(a)
variance, an applicant must demonstrate
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that effluent controls on its thermal
discharge will be sufficient to assure the
protection and propagation of a
balanced indigenous population of
shellfish, fish and wildlife in and on the
water. The Agency agrees that in most
cases the demonstration required of a
successful section 316(a) applicant will
be sufficient to allow the permit writer
to conclude that there will be no
unreasonable degradation of the marine
environment due to excess heat. While
on the reasoning set out in the previous
response successful section 316(a)
applicants will not be exempt from
section 403, the regulation provides that
a successful section 316(a) application
creates a rebuttable presumption of
compliance with section 403(c) for the
thermal component of the discharge.
11. Comment. Commenters also
suggested that those publicly-owned
treatment works which installed
secondary treatment should be exempt
from requirements under section 403.
Response. Limitations established
pursuant to section 403 are a supplement
to technology-based limitations such as
secondary treatment for POTWs, and no
class of a discharger is exempt from
compliance with these regulations.
However, it is likely that secondary
treatment will generally be adequate to
satisfy section 403 requirements where
there is adequate pretreatment by
industrial sources and whers the POTW
is not discharging into areas of
biological concern.
12. Comment. As noted above, some
commenters asserted that the ocean
discharge criteria should merely be
guidelines providing information to the
permit writer. Other commenters,
however, stated that the guidelines
should require that a discharge pass a
quantitative test, such as the bioassay
requirements used in the ocean dumping
regulations, and comply with State and
EPA water quality criteria as a
prerequisite to permit issuance.
Response. The Agency has revised the
proposed regulation to allow necessary
flexibility to the director in assessing
both the impact of a discharge and
permit limitations. However, the
regulation does impose minimum permit
limitations, including a bioassay-based
limitations, in areas where the long
range impact of a discharge is not fully
understood. This approach should
provide certainty and consistency in
permit limitations in areas where the
determinations by the permit writer
would be the most difficult and complex.
Discharges into the territorial seas
must comply with any applicable state
water quality criteria. However, the Act
generally does not provide for the
application of these criteria to the
contiguous zone and oceans. Although
the Act establishes a complete water
quality program for State waters based
on designated uses and supporting
criteria, no such scheme exists for
marine waters beyond State jurisdiction.
The 403(c) regulation is consistent with
the Agency policy outlined in the section
301(h) regulations (44 FR 34810-34811),
and will utilize water quality criteria
published pursuant to section 304{a)(l),
as they are developed, as a basis for
assessing the environmental impact of
such pollutants.
13. Comment. Several commenters
asserted that, under the proposed
regulation, no predischarge
determination was required by the
director to assure that the marine
environment was protected. Instead,
commenters stated, the proposed
regulation relied on post discharge
monitoring.
Response. Under the final regulations,
no discharge of pollutants may be
authorized unless, before permit
issuance, the director has sufficient
information to make a reasonable
determination that there will be no
irreparable harm to the environment
while monitoring is undertaken to
determine if there will be unreasonable
degradation. In addition the permit must
specify certain mandatory limitations.
14. Comment. The Agency received
numerous comments regarding the
monitoring requirements outlined in
section 125.127 of the proposed
regulation. The major issue raised was
that the monitoring requirements should
be as flexible as possible providing an
applicant a clear description of the
information he must provide.
Commenters suggested that the rigor ofi
the monitoring program should be
tailored to site-specific conditions such
as the nature and location of the
discharge. In addition, a number of
commenters stated that compliance with
the proposed monitoring requirements
would result in severe economic
hardship for small dischargers. It was
suggested that the latter, especially
small POTWs, be exempted from the
monitoring requirements.
Response. As discussed in the
preamble, these regulations have been
revised to allow the permit writer to
request from the applicant only that
information necessary to make
judgments required by the guidelines. In
some cases this will involve monitoring
programs, and the director will work
with the applicant in identifying specific
information that must be supplied as
part of the permit application process.
Since no discharge may be allowed
which would result in unreasonable
degradation of the marine environment,
and since the permit writer must be
afforded the means to make the
necessary determinations under these
regulations, EPA has concluded that it
may not exempt categories of
dischargers, even small dischargers,
from monitoring requirements as an
initial matter. Nevertheless, the final
regulations do not require monitoring in
all cases, and, where monitoring is
necessary, provide for flexibility in
fashioning site specific requirements.
Although any monitoring that may be
necessary will depend on the nature and
location of the discharge in question.
small dischargers generally are not
expected to incur significant economic
costs as a result of this regulation.
15. Comment. Several commenters
suggested that, in light of the similarities
between section 403(c) of the Clean
Water Act and section 102(a) of the
Marine Protection, Research and
Sanctuaries Act, the. ocean discharge
criteria should be similar to the ocean
dumping regulations.
Response. EPA recognizes that in
section 403(c) of the Clean Water Act
and section 102(a) of the Marine
Protection. Research, and Sanctuaries
Act, Congress adopted similar although
not identical provisions. Hence, in the
regulations implementing the respective
statutes, similar criteria may be
appropriate.
EPA first promulgated ocean dumping
criteria in 1973; those criteria were
amended in 1977. Initially, the
regulations served as joint regulations
for the CWA and the MPRSA. Since
promulgation of the ocean dumping
regulations, however, EPA has received
a number of comments based on those
regulations. In addition, increasing data
has become available in respect to the
environmental impact of disposing of
material at various locations in the
ocean, by various methods.
The ocean discharge regulations being
promulgated today are based on the
latest data and information available to
EPA, and the Agency believes these
regulations are consistent with the CWA
and with current scientific and technical
knowledge. Various factors, including
the MPRSA comments and the new
data, suggest that it may now be
appropriate for EPA to review the ocean
dumping regulations as well. Such a
review may provide further insights on :
an appropriate overall approach for
protecting the ocean; and
inconsistencies which may exist
between the current sets of regulations
can be resolved in the context of that
action. However, in addition to any
statutory distinctions, differences in uw
manner of disposal and the types of .
pollutants discharged may warrant
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Federal Register / Vol. 45. No. 194 / Friday October 3, 1980 / Rules and Regulations
65953
Different regulatory approaches under
jhese two statutes.
16, Comment. Several commenters
niegested that dischargers who were not
causing unreasonable degradation of the
nanne environment should not be
-quired to assess the availability of
alternatives. _
Response. The proposed regulation
jjt» been modified to require the
usesjment of reasonable alternatives
only where the director cannot
(Jeterraine whether the discharge will
cat^e unreasonable degradation of the
jjjrine environment. In cases where the
director determines that the discharge
will not cause such degradation, an
NPDES permit may be issued
notwithstanding the availability of an
illemative to ocean disposal.
Although the Clean Water Act
contains as an ultimate goal the
complete elimination of the discharge of
pollutants, the water quality provisions
of Ihe Act. including sections 303 and
403, dq not require that discharges into
either inland or offshore waters be
prohibited in the absence of
unreasonable water quality impacts.
. 17. Comment. Several commenters
expressed concern that the requirements
{or general permits, specified in section
125.129 of the proposed regulation, were
aot consistent with the requirements fof
Individual permits*.
Response. Section 125.129 has been
deleted from the final regulation, and
lt« director is required to make the
uzne determinations when issuing
either general or individual permits.
18. Comment. A few commenters
objected to language in the preamble of
&e proposal to the effect that the
J*raiiUing authority would be free to
&aw on his own knowledge of
conditions in the vicinity of an outfall in
Determining whether a discharge
xJversely affected the marine
*nvironment. These commenters
expressed the "due process" concern
"•' this language allowed the
nitling authority to issue permits on
basis of information not made
to the permittee and not in the
Mministrative record.
e. The language in question
**« not appear in the final regulation or
nits preamble. While the regulation
Provides that the director make the
"tlerminatlons under §.125.123(a), [b),
? w) on the basis of "available
^formation." that language was added
* response to the suggestion of
oamerous commenters that permit
*PP!icants should not be required for
P^Poses of Section 403 to resubmit data
"fitch was already available to the
•""nit writer.
19. Comment. EPA received comment
to the effect that the reference in the •
proposed regulations to schedules
allowing additional time for compliance
with Section 403'requirements should be
limited to existing dischargers,
consistent with the provisions of the
NPDES regulations.
Response. Section 125.123(d)(3) of the
final regulation provides for "schedules
of compliance for existing dischargers,"
as suggested above.
20. Comment. The Agency received
several comments regarding the mixing
zone analysis as described in § 125.123.
Some commenters suggested that the
models identified by the Agency in the
proposal would not be appropriate for
all types of discharge. Other
commenters suggested that the Agency
should use the ocean dumping mixing
zone definition.
Response. The mixing zone analysis
in the final regulations is intended for
use in calculating whether the limiting
permissible concentration is violated in
instances where bioassay analysis is
required. The proposed regulation
required a mixing zone analysis for all
.dischargers to assure that, following
initial dilution, the discharge was
dispersed so as not to adversely affect
areas of biological sensitivity. As noted
previously, this requirement has been
deleted from the final regulation. In the
final regulation, a mixing zone analysis
is required only in those instances
where the director cannot determine
that unreasonable degradation will not
occur and where a bioassay analysis is
required.
The mixing zone definition in these
regulations is consistent with the ocean
dumping mixing zone definition
identified in the EPA/Corps of Engineers
technical manual, with some
modifications to account for the
differences in the nature of discharged
wastes versus those which are dumped.
The ocean dumping mixing zone was
devised primarily to facilitate analysis
of impacts from intermittent discharges
from moving vessels, whereas the 403(c)
regulations are intended to facilitate
analysis of continuous discharges from
stationary sources. The final regulation
also allows the discharger to use
alternative methods for determining the
mixing zone where scientific evidence
demonstrates they are appropriate and
where EPA concurs.
A new Subpart M is added to read as
follows:
Subpart M—Ocean Discharge Criteria
Sec.
125.120 Scope and purpose. '
125.121 Definitions.
Sec.
125.122 Determination of unreasonable
degradation of the marine environment.
125.123 Permit requirements.
125.124 Information required to be
submitted by applicant.
§ 125.120 Scope and purpose.
This subpart establishes guidelines for
issuance of National Pollutant Discharge
Elimination System (NPDES) permits for
the discharge of pollutants from a point
source into the territorial seas, the
contiguous zone, and the oceans.
§ 125.121 Definitions.
(a) "Irreparable harm" means
significant undesirable effects occurring
after the date of permit issuance which
will not be reversed after cessation or
modification of the discharge.
(b) "Marie environment" means that
territorial seas, the contiguous zone and
the oceans.
(c) "Mixing zone" means the zone
extending from the sea's surface to
seabed and extending laterally to a
distance of 100 meters in all directions
from the discharge point(s) or to the
boundary of the zone of initial dilution
as calculated by a plume model
approved by the director, whichever is
greater, unless the director determines
that the more restrictive mixing zone or
another definition of the mixing zone is
more appropriate for a specific
discharge.
(d) "No reasonable alternatives"
means: (1) No land-based disposal sites,
discharge point(s) within internal
waters; or approved ocean dumping
sites within a reasonable distance of the
site of the proposed discharge the use of
which would not cause unwarranted
economic impacts on the discharger, or,
notwithstanding the availability of such
sites,
(2) On-site disposal is
environmentally preferable to other
alternative means of disposal after
consideration of: (i) The relative
environmental harm of disposal on-site,
in disposal sites located on land, from
discharge point(s) within internal
waters, or in approved ocean dumping
sites, and
(ii) The risk to the environment and
human safety posed by the
transportation of the pollutants.
(e) "Unreasonable degradation of the
marine environment" means: (1)
Significant adverse changes in
ecosystem diversity, productivity and
stability of the biological community
within the area of discharge and
surrounding biological communities,
(2) Threat to human health through
direct exposure to pollutants or through
consumption of exposed aquatic
organisms, or
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Appendix B
Federal Register / Vol. 45. No. 194 / Friday, October 3. 1980 / Rules and Regulations
(3) Loss of esthetic, recreational,
scientific or economic values which is
unreasonable in relation to the benefit
derived from the discharge.
§ 125.122 Determination of unreasonable
degradation of the marine environment.
(a) The director shall determine
whether a discharge will cause
unreasonable degradation of the marine
environment based on consideration of:
(1) The quantities, composition and
potential for bioaccumulation or
persistence of the pollutants to be
discharged;
(2) The potential transport of such
pollutants by biological, physical or
chemical processes;
(3) The composition and vulnerability
of the biological communities which
may be exposed to such pollutants,
including the presence of unique species
or.communities of species, the presence
of species identified as endangered or
threatened pursuant to the Endangered
Species Act, or the presence of those
species critical to the structure or
function of the ecosystem, such as those
important for the food chain;
(4) The importance of the receiving
water area to the surrounding biological
community, including the presence of
spawning sites, nursery/forage areas,
migratory pathways, or areas necessary
for other functions or critical stages in
the life cycle of an organism.
(5) The existence of special aquatic
sites including, but not limited to marine
sanctuaries and refuges, parks, national
and historic monuments, national
seashores, wilderness areas and coral
reefs;
(6) The potential impacts on human
health through direct and indirect
pathways;
(7) Existing or potential recreational
and commercial fishing, including
finishing and shellfishing;
(8) Any applicable requirements of an
approved Coastal Zone Management
plan;
(9) Such other factors relating to the
effects of the discharge as may be
appropriate;
(10) Marine water quality criteria
developed pursuant to section 304(a)(l).
[b) Discharges in compliance with
sections 301(g), 301(h), or 316(a)
variance requirements or State water
quality standards shall be presumed not
to cause unreasonable degradation of
the marine environment, for any specific
pollutants or conditions specified in the
variance or the standard.
§ 125.123 Permit requirements.
(a) If the director on the basis of
available information including that
supplied by the applicant pursuant to
§ 125.124 determines prior to permit
issuance that the discharge will not
cause unreasonable degradation of the
marine environment after application of
any necessary conditions specified in
§ 125.123(d), he may issue an NPDES
permit containing such conditions.
(b) If the director, on the basis of
available information including that
supplied by the applicant pursuant to
§ 125.124 determines prior to permit
issuance that the discharge will cause
unreasonable degradation of the marine
environment after application of all
possible permit conditions specified in
§ 125.123(d), he may not issue an NPDES
permit which authorizes the discharge of
pollutants.
(c) If the director has insufficient
information to determine prior to permit
issuance that there will be no
unreasonable degradation of the marine
environment pursuant to § 125.122, there
shall be no discharge of pollutants into
the marine environment unless the
director on the basis of available
information, including that supplied by
the applicant pursuant to § 125.124
determines that: (1) Such discharge will
not cause irreparable harm to the
marine environment during the period in
which monitoring is undertaken, and
(2) There are no reasonable
alternatives to the on-site disposal of
these materials, and
(3) The discharge will be in
compliance with all permit conditions
established pursuant to paragraph (d) of
this section.
(d) All permits which authorize the
discharge of pollutants pursuant to
paragraph (c) of this section shall: (1)
Require that a discharge of pollutants
will: (A) following dilution as measured
at the boundary of the mixing zone not
exceed the limiting permissible
concentration for the liquid and
suspended particulate phases of the
waste material as described in section
227.27(a) (2) and (3), section 227.27(b),
and section 227.27(c) of the Ocean
Dumping Criteria; and (B) not exceed
the limiting permissible concentration
for the solid phase of the waste material
or cause an accumulation of toxic
materials in the human food chain as
described in sections 227.27 (bj and (d)
of the Ocean Dumping Criteria;
(2) Specify a monitoring program,
which is sufficient to assess the impact
of the discharge on water, sediment, and
biological quality including, where
appropriate, analysis of the
bioaccumulative and/or persistent
impact on aquatic life of the discharge;
(3) Contain any other conditions, such
as performance of liquid or suspended
particulate phase bioaccumulation tests,
seasonal restrictions on discharge.
process modifications, dispersion of
pollutants, or schedule of compliance for
existing discharges, which are
determined to be necessary because of
local environmental conditions, and
(4) Contain the following clause: In
addition to any other grounds specified
herein, this permit shall be modified or
revoked at any time if, on the basis of
any new data, the director determines
that continued discharges may cause
unreasonable degradation of the marine
environment.
§ 125.124 Information required to be
submitted by applicant
The applicant is responsible for
providing information which the director
may request to make the determination
required by this subpart. The director
may require the following information as
well as any other pertinent informaton:
(a) An analysis of the chemical
constituents of any discharge;
(b) Appropriate bioassays necessary
to determine the limiting permissible
concentrations for the discharge;
{c) An analysis of initial dilution;
(d) Available process modifications
which will reduce the quantities of
pollutants which will be discharged;
(e) Analysis of the location where
pollutants are sought to be discharged,
including the biological community and
the physical description of the discharge
facility;
(f) Evaluation of available alternatives
to the discharge of the pollutants
including an evaluation of the possibility
of land-based disposal or disposal in an
approved ocean dumping site.
[FR Doc. 80-30723 Filed 10-2-80:8:45 am)
BILLING CODE 6560-O1-M
U.S. GOVERNMENT PRINTING OFFICE: 1994 — 5
15-003 /01006
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