TD427
.M44I58
1992
United States
Environmental Protection
Agency
Office of Science and Technology
Health and Ecological Criteria Division
Washington, DC 20460
May 1992
OOOR92003
Interim Guidance on
Interpretation and Implementation
of Aquatic Life Criteria for Metals
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Interim Guidance on
Interpretation and Implementation of
Aquatic Life Criteria for Metals
May 1992
U.S. environmental Protection Agency
RPw5.Ubrary (PL-12J)
West Jackson Boulevard, 12th Ffonr
ic, IL 60604-3590 °r
Health and Ecological Criteria Division
Office of Science and Technology
U.S. Environmental Protection Agency
Washington, DC 20460
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This document provides guidance for interpreting and im-
plementing aquatic life criteria for metals in waters of the
United States. It is issued in support of EPA regulations and
initiatives involving the application of water quality criteria and
standards. This document is agency guidance only. It does not es-
tablish or affect legal rights or obligations. It does not establish a
binding norm, or prohibit alternatives not included in the docu-
ment. It is not finally determinative of the issues addressed.
Agency decisions in any particular case will be made by applying
the law and regulations on the basis of specific facts when regula-
tions are promulgated or permits are issued.
This document is expected to be revised periodically to reflect
advances in this rapidly evolving area. Comments from readers
are welcomed. Send comments to Health and Ecological Criteria
Division (WH-586), U.S. EPA, 401 M Street SW, Washington, DC
20460.
Foreword
iii
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This guidance addresses the use of EPA (and corresponding
State) metals criteria in water quality standards intended to
protect aquatic life. This guidance also addresses the deriva-
tion of NPDES permit limits from such criteria. The main body of
the document presents recommendations on the best current ap-
proaches for implementing aquatic life criteria for metals and
measuring attainment of such criteria. This guidance supersedes
past criteria document statements expressing criteria in terms of
an acid soluble analytical method. Appendix A presents a case
study illustrating derivation of site-specific criteria (item 3 below).
Appendix B presents recommendations on the derivation of
NPDES permit limits from ambient metals criteria. As described in
Appendix B, it supersedes part of the Technical Support Docu-
ment [1] discussion of metals.
The principal issue is the correlation between metals that are
measured and metals that are biologically available. The
bioavailability and toxicity of metals depend strongly on the exact
physical and chemical form of the metal, and on the species af-
fected. The form of the metal, in turn, can vary depending on the
chemical characteristics of the surrounding water matrix. Because
of differences between various effluents and site waters, and be-
tween laboratory toxicity test waters and many site waters,
establishment and implementation of metals criteria are not
straight forward. Consequently, this guidance presents three
reasonable approaches that differ in their complexity.
(1) The simplest approach is to measure total recoverable
metals in ambient waters, and to compare such meas-
urements to national or state-wide criteria.
(2) A closer focus on biologically available metals can be
obtained by measuring dissolved metals in ambient
waters, and comparing such measurements to criteria
appropriate for dissolved metal. Since effluent limits,
for both technical and legal (40 CFR 122.45) reasons,
are generally expressed in terms of total recoverable
metal, it is necessary to translate between the total
recoverable concentration in the effluent and the dis-
solved concentration in the ambient water.
(3) Because of the complexity of metal chemistry, there is
no one chemical analytical method that can accurately
determine the metals that are bioavailable and toxic.
For implementing metals criteria established from
laboratory toxicity tests, an adjustment of the criteria
value can address this constraint. It involves measur-
Synopsis
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Introduction
ing a pollutant's water-effect ratio in the receiving
water covered by the standard. The water-effect ratio
compares the toxicity of a pollutant in the actual site
water to its toxicity in laboratory water, for two or
more aquatic species. Because the metal's toxicity in
laboratory water is the basis for the national criterion,
the water-effect ratio is used in an adjustment to ob-
tain a site-specific value. Implemented in conjunction
with either of the first two alternatives, this adjust-
ment may either increase or decrease the numeric
value of the criterion.
The principal problem in relating discharges of toxic metals to
environmental impacts is the different toxicities of various
metal species in ambient waters, and the varying fractions of
such species with location and time. This results in the same metal
concentration exerting different toxicity from place to place and
from time to time. The chemical species involved include metals
dissolved in a variety of forms, and metals sorbed to or within
particulate matter. Metals may differ markedly from each other
with respect to speciation and bioavailability.
Although metal toxicity may vary depending on the chemical
characteristics of the water body, the national criteria have been
designed to protect all or almost all bodies of water. However, this
does not mean that the national criteria will always be overprotec-
tive. For example, some untested locally important species might
be very sensitive to the material of concern, or the local aquatic or-
ganisms might have increased sensitivity due to diseases,
parasites, other pollutants or water quality conditions, or extreme
flow or temperature conditions [2].
Another problem involves metal speciation in effluents, and
the potential transformations that may occur in moving from the
chemical environment of the effluent to the chemical environment
of the receiving water. Consequently, in contrast to an ambient
measurement, which should respond predictably to metal that is
actually bioavailable, an effluent measurement needs to respond
also to metal that may not be bioavailable under effluent chemical
conditions, but would possibly become bioavailable under am-
bient chemical conditions.
Because of the complexity of metal speciation and its effect on
toxicity, the relationship between measured concentrations and
toxicity is not precise. Consequently, any chemical analytical
method that could be recommended would not guarantee precise
comparability between concentrations measured in the field and
concentrations employed in the toxicity tests underlying the
criteria. However, the three approaches presented in this guidance
should provide acceptable approximations.
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EPA has recognized four methods of sample preparation for
metals analysis. These lead to measurement of: (a) total me-
tals, (b) total recoverable metals, (c) acid soluble metals, and
(d) dissolved metals. Ordinarily, the four methods measure all of
the dissolved metal present at the time of sampling. They differ in
the amount of particulate metal that they measure.
The total metals procedure, the total recoverable metals proce-
dure [3], and the acid soluble metals procedure [4, 5] measure
metals that are dissolved in water or become dissolved when
treated with acid. They differ in the concentration of acid and in
the temperature used during the analytical procedure, both
decreasing in the order cited above.
The dissolved procedure [3] measures metal that passes
through a 0.45 ^m filter at the time of sample collection. The
results from this procedure are reported as "dissolved," although
it may include metal that was bound to micro-particles (<0.45 ^im)
at the time of sample collection. More recent dissolved procedures
recommend positive-pressure, in-line filtration through poly-
carbonate membrane filters having a uniform pore size selected
from a range of 0.1-0.4 ^m [6], and emphasize ultra-clean
laboratories, labware, and reagents [7, 13]. Measurements using
different filter sizes may, however, give different results.
Metals criteria documents issued in 1980 recommended the
use of the total recoverable method. Beginning in 1984, although a
final acid soluble method was not available, the criteria docu-
ments have stated that an acid soluble method would be a better
way of measuring attainment of the criteria. Noting the un-
availability of a final method, they recommended the continued
use of the total recoverable method, which they acknowledge may
be overly protective.
Because the acid soluble method uses a less rigorous digestion,
it was expected that it would generally measure less of the particu-
late metal than the total recoverable method. It was therefore
believed that the acid soluble method would more accurately
measure bioavailable metal. Recently available ambient and ef-
fluent data suggest, however, that acid soluble results are
ordinarily nearly identical to total recoverable results, while being
somewhat different from dissolved results. Because an increased
understanding of the complexity of metals bioavailability indi-
cates that the acid soluble method will not significantly improve
the correlation between measured metal and bioavailable metal,
this guidance is not recommending the use of this method.
Bioavailability and toxicity vary with the form of the metal.
Particulate metal is generally expected to have less
bioavailability than dissolved metal. Nevertheless, the
toxicity of ambient particulate metal is not necessarily zero. For
example, some metal that is in the particulate phase in the ambient
Background
on Analytical
Methods
Bioavailability
and Toxicity
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Dissolved
and Total
Recoverable
Approaches
water environment may become dissolved in the chemical en-
vironment associated with the gill or the gut.
In natural waters, some metals may exist in a variety of dis-
solved species that differ significantly in toxicity. For copper, the
divalent free cation and some inorganic complexes have substan-
tial toxicity, whereas dissolved organic complexes generally have
significantly less toxicity. As a result, the same concentration of
dissolved copper may exert different toxicity in different waters.
Toxicity tests that form the basis for the criteria are usually per-
formed in an untreated or slightly treated natural water from an
uncontaminated source, or in water that has been first purified
and then reconstituted by the addition of appropriate mineral
salts. Because such dilution water is generally lower in metal-
binding particulate matter and dissolved organic matter than most
ambient waters, these toxicity tests may overstate the ambient
toxicity of non-biomagnified metals that interact with particulate
matter or dissolved organic matter.
In most but not all toxicity tests underlying the criteria, the
percentage of metal in the particulate phase is fairly low. For am-
bient waters, on the other hand, recent data suggest that typically
30-80 percent of the copper, nickel, and zinc, and 90-95 percent of
the lead may be in a particulate phase measured by the total
recoverable method but not by the dissolved method.
In freshwater laboratory tests, organic carbon concentrations
of a few mg/L are typical, with chronic tests having higher con-
centrations than acute tests. In ambient waters, organic carbon
concentrations are typically somewhat higher than this, and may
be substantially higher at the edge of small mixing zones.
Because of the greater fraction of particulate metal in ambient
waters, as well as the higher levels of dissolved organic binding
agents in ambient waters, the fraction of metal that is biologically
available may often be lower under ambient field conditions than
under laboratory conditions, particularly for freshwaters.
Aquatic life criteria in ambient waters may be implemented
either as total recoverable metal or as dissolved metal. Ef-
fluent limits must generally be expressed as total
recoverable metal. For analyses of metals in the low ng/L range or
below, ultra-clean sample handling techniques [7, 13] should be
used.
Ambient Waters
When used for expressing ambient water quality criteria, the total
recoverable method provides greater safety than does the
dissolved method. Nevertheless, when used for ambient waters,
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total recoverable measurements may result in overestimating the
toxicity. While toxicity testing has shown dissolved measurements
to be better predictors of toxicity than total recoverable
measurements, there are also some potential concerns with this
approach, as discussed below.
First, EPA water quality criteria are generally based on the
reported total recoverable concentrations in the toxicity tests. If
used for dissolved standards, the criteria values need to be
downwardly adjusted to account for the typical dissolved fraction
in test dilution water. For copper, approximately 86 percent of the
reported total concentration was dissolved during freshwater
acute toxicity tests with the more sensitive species. Consequently,
the copper freshwater acute criterion should be adjusted to 86 per-
cent of its total recoverable criterion in order to serve as a
dissolved criterion, particularly in waters having low concentra-
tions of dissolved organic binding agents. While the adjustment
may be small for most metals, a few metals, such as aluminum,
may require much larger adjustments to account for the much
lower percentage dissolved typically occurring during toxicity
tests. Chronic criteria may require larger adjustments than acute
criteria, due to the higher particulate concentrations caused by the
addition of food during chronic tests.
Except for copper in freshwater, the factors are not yet avail-
able for converting EPA's published criteria into dissolved criteria.
A re-examination of data underlying the metals criteria is now un-
derway to compile the dissolved concentrations measured during
toxicity tests. While preliminary analysis does not indicate that
these dissolved adjustment factors are of sufficient magnitude to
be of great concern, they should be considered in any adoption of
dissolved metal standards subsequent to distribution of this infor-
mation.
Second, by measuring comparatively little of the particulate
fraction, it may be possible that the dissolved method could oc-
casionally understate the toxicologically effective concentration.
Although toxicity data suggest that this is not ordinarily a prob-
lem, it is more likely to be a concern if the dissolved concentration
is only a very small percentage of the particulate concentration,
such as may occur with aluminum.
In some situations the dissolved method may overstate the
toxicologically effective concentration. When certain metals (such
as copper) become complexed with elevated concentrations of dis-
solved organic matter, a reduction in toxicity may occur, compared
to toxicity in laboratory water, which is low in organic matter.
Where dissolved organic matter is likely to interact with the
toxicant, the water-effect ratio approach is likely to be more ac-
curate and is currently the recommended solution.
A review of the limited number of available site-specific
studies found that the water-effect ratio (site water LC50 versus
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lab water LC50) was generally significantly larger than the
measured total recoverable versus dissolved ratio [10]. These
limited freshwater data thus suggest that use of properly formu-
lated dissolved criteria would be at least as protective as criteria
derived from careful measurements of water-effect ratios.
Effluents
The dissolved method is generally not applied to effluents to
determine achievement of effluent quality goals. Such use is
generally barred by regulation (40 CFR 122.45). Because the
chemical conditions in ambient surface waters may differ sub-
stantially from those in the effluent, there is no assurance that
effluent particulate metal would not dissolve after discharge.
A common method of removing metals from wastewaters is to
chemically precipitate the metal and settle the resulting particles.
Expressing a metals limitation in terms of dissolved metal would
imply little concern about the effectiveness of the settling process
or the fate of the discharged particulate metal.
Determining the total recoverable effluent limitation cor-
responding to a dissolved criterion would involve specifying the
fraction of effluent total recoverable metal that would exist as dis-
solved metal under the chemical conditions of the receiving water.
In the absence of site information, any values assumed for this
fraction should be environmentally conservative.
Where greater accuracy is desired, the dissolved fraction of
total recoverable metal could be evaluated by direct measurement
of dissolved and total recoverable metal in the affected ambient
waters, or possibly by geochemical model calculations (as dis-
cussed in Appendix B). All of the techniques involve approx-
imations.
Water-Effect
Ratio
Approach
Due to the complexity of metals speciation, and due to the
varying degrees of bioavailability and toxicity of the many
forms and complexes, there is no chemical method that can
assure that a unit of concentration measured in the field would al-
ways be lexicologically equivalent to a unit of concentration
employed in the laboratory toxicity tests underlying the criteria.
For metals criteria derived from laboratory toxicity tests, one
approach is to use a biological method to compare bioavailability
and toxicity in receiving waters versus laboratory test waters. This
involves running toxicity tests with at least two species, measur-
ing acute (and possibly chronic) toxicity values for the pollutant
using (a) the local receiving water, and (b) laboratory toxicity test-
ing water, as the sources of toxicity test dilution water. A
water-effect ratio is the acute (or chronic) value in site water
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divided by the acute (or chronic) value in laboratory waters. An
acute value is an LC50 or EC50 from a 48-96 hour test, as ap-
propriate for the species. A chronic value is a concentration
resulting from hypothesis testing or regression analysis of meas-
urements of survival, growth, or reproduction in life cycle, partial
life cycle, or early life stage tests with aquatic species.
Because the metal's toxicity in laboratory water is the basis for
EPA's criterion, this water-effect ratio is used to adjust the national
criterion (or corresponding State criterion) to a site-specific value.
This adjustment may either increase or decrease the criterion. Be-
cause the water-effect ratio reflects differences in water chemistry,
it is acceptable to assume that a ratio derived from acute LCSOs or
ECSOs may be applied to both acute and chronic criteria, provided
that the water-effect ratio is determined with an acutely sensitive
species. Nevertheless, performing chronic tests is an option that
could produce a different water-effect ratio, due to changes in
water chemistry caused by the addition of food during chronic
tests. While this may somewhat improve the accuracy of the
resulting criteria, it will substantially increase the testing costs.
The principle of criteria adjustment using a water-effect ratio
was set forth in previous guidance [8, 9]. The procedure applies to
criteria derived from laboratory toxicity data. As such, it does not
apply to the residue-based mercury chronic criteria, or the field-
based selenium freshwater criterion. The basic features of the
procedure as recommended herein are described below. Dis-
chargers or private entities wishing to perform such testing should
consult with the appropriate regulatory agency before proceeding.
(1) At least two sensitive species, including at least one in-
vertebrate, should be tested through standard toxicity
testing protocols, using site dilution water and using
laboratory dilution water. Test organisms should be
drawn from the same population and tested under
identical conditions (except for water source). Test
species should ordinarily be selected from those
species that were used for national criteria develop-
ment in order to be able to ascertain whether the
laboratory water results are comparable to the those in
the criteria document.
(2) Site water samples used for testing are to be repre-
sentative of the receiving water to which the site-
specific criteria value is to apply. For flowing waters, it
is generally recommended that at least one sample
correspond to a condition in which point or nonpoint
pollutant contributions are reasonably well mixed
with the flow of the receiving water. For other types of
waters, it is generally recommended that a sample cor-
respond to a dilution situation well outside any
regulatory mixing zone. These recommendations are
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intended to yield a water-effect ratio appropriate for
the affected receiving water as a whole. These recom-
mendations supersede those of the previous site-
specific guidance [8, 9], which recommended that
pristine waters always be used.
(3) The laboratory dilution water should be comparable
to what was used in tests underlying the national
criteria. For any pollutant with a national or State
criterion calculated from site-specific hardness,
laboratory-water and site-water toxicity results should
be computationally normalized to the same hardness,
using the specified hardness slope.
(4) The toxic metal should be added in the form of an in-
organic salt having relatively high solubility. Nitrate
salts are generally acceptable; chloride and sulfate
salts of many metals are also acceptable. Results
should be based on measured or nominal initial con-
centrations if static tests are performed, and on
average measured concentrations if flow-through tests
are performed.
(5) Water quality characteristics affecting bioavailability
and toxicity should be monitored. Measurements or-
dinarily should include both dissolved and total
recoverable metal concentrations, hardness, pH,
alkalinity, suspended solids, conductivity, dissolved
solids or salinity, total organic carbon, dissolved or-
ganic carbon, temperature, and specific metal binding
ligands (where known to be important).
(6) The number of site water samples to be tested may
vary with the size of the affected water body (or the
size of the metal loading). Except in the smallest sys-
tems, a minimum of two site water samples should be
collected in different seasons during times of relatively
low flow or low dilution. In moderately sized and
larger systems (e.g., multiple m3/sec or double to
triple digit cfs low flow), additional samples should be
collected during other times in the year and possibly
at additional locations appropriate for the segment
under study.
(7) In studies involving continuous discharges, samples
ordinarily should not be taken from storm affected
waters, which may contain particulate matter not
present during design flow conditions. On the other
hand, in effluent dominated situations, at least one
sample should represent a higher dilution condition
(less than 50 percent effluent) in order to portray am-
bient conditions. In all situations, it should be recog-
8
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nized that the water-effect ratio may be affected by
constituents contributed by point and nonpoint
sources. Consequently, new measurements should be
undertaken if newly implemented controls or other
changes substantially affect ambient levels of
suspended solids, organic carbon, or pH.
(8) Additional testing should be performed before accept-
ing unusually or inexplicably high values for the
water-effect ratio, based on experience with the par-
ticular pollutant, and based on the chemical charac-
teristics of the water. Retesting should also be
performed for ratios having wide uncertainty ranges.
EPA intends to compile additional information to as-
sist in judging water-effect ratios in this manner. These
recommendations, which focus concern on large and
uncertain water-effect ratios, supersede the previous
guidance [8, 9] recommendations that encourage
retesting or rejection of small water-effect ratios.
(9) Ordinarily, the acute and chronic criteria for the site
are calculated by multiplying the national or State
criteria by the geometric mean water-effect ratio for
the two or more tested species. The previous site-
specific guidance [8, 9] provides some additional sug-
gestions for situations where the measured ratios
differ significantly, and provides other alternatives for
setting the chronic criterion.
(10) As with other types of water quality-based control ac-
tions for toxic pollutants, it is recommended that the
chemical-specific approach be implemented in con-
junction with assessments of whole effluent toxicity
and field ecology ("bioassessment") [1]. Nevertheless,
in light of the stated limitations of these latter techni-
ques with regard to identifying causative agents and
predicting future changes [1], considerable caution is
warranted in using such information (particularly
ecological data) to make inferences about the ade-
quacy of particular numeric criteria.
The water-effect ratio is affected not only by speciation among
the various dissolved and particulate forms, but also by additive,
synergistic, and antagonistic effects of other materials in the af-
fected site waters. As such, the water-effect ratio is a much more
comprehensive measure than a ratio of total recoverable metal to
dissolved metal. Because the basic technique involves adding
soluble metal salts to site water samples, it is most accurate where
rapid sorption or complexation processes are involved.
Because effluent limits are generally expressed as total
recoverable metal, simplicity would suggest deriving water-effect
ratios in terms of total recoverable measurements. Derivation in
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Relationship
with Sediment
Criteria
Current
Technical
Support and
Future
Research
terms of dissolved measurements is also acceptable, and may be
preferred in situations involving highly variable suspended solids
concentrations.
Data available from a limited number of site-specific studies
performed in rather clean freshwater suggest that copper, lead,
and cadmium often have substantial water-effect ratios, while
zinc, in situations where it is preponderantly dissolved, often does
not [10]. Much less information is available for such ratios in salt
water.
For national or state-wide criteria expressed as dissolved or
total recoverable metal, and for site-specific criteria derived
from water-effect ratios, questions may be raised about the
adequacy of water column criteria for protecting sediment. The
issue is whether particulate metal settling from the water column
could contribute to sediment quality problems, even where no
toxicity is manifested in the water column.
Because sediment toxicity is considered to be determined
primarily by the concentrations of pollutant dissolved in the sedi-
ment interstitial water, the question becomes whether the
pollutant would have a greater propensity to become dissolved or
bioavailable in the sediment than in the water column. While
available information does not suggest that this is ordinarily the
case, the ongoing development of sediment criteria should resolve
this issue. Nevertheless, in those cases where the beneficial uses of
a receiving water are known to be impaired by the toxicity of me-
tals in sediments, water quality-based control requirements
should be designed to abate any sources that would continue to
cause sediment toxicity.
The Environmental Research Laboratories in Duluth and in
Narragansett will continue to answer technical questions
about the possible problems in applying the above methods
to criteria for specific metals. The contact for freshwater is Charles
Stephan (Duluth, Minnesota telephone (218) 720-5510). The con-
tact for salt water is Gary Chapman (Newport, Oregon telephone
(503) 867-4027).
EPA intends to undertake further work to facilitate the im-
plementation of metals criteria in terms of dissolved
measurements. For metals such as copper, silver, zinc, lead, and
cadmium, the dependency of toxicity on factors other than hard-
ness will be considered for inclusion. Where appropriate and
feasible, EPA may develop equations relating dissolved metal
criteria to hardness and organic matter concentration, and possib-
ly pH and other water quality characteristics. EPA might also
consider other biological or chemical techniques for ascertaining
the effective concentration of bioavailable metals.
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APPENDIX A
CASE STUDY:
Determination of the Water-Effect Ratio
Using Indicator Species
Norwalk River
Georgetown, Connecticut*
* Adapted from the 1983 Water Quality Standards Handbook [8] for
purposes of illustrating an application.
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Connecticut's Upper Norwalk watershed, where this study
was conducted, covers an area of 18.5 square miles and in-
cludes the region extending from the headwaters of the
Norwalk River to its confluence with Comstock Brook.
Two secondary treatment plants discharge a total of 0.44 mgd
of municipal wastewater to a reach 9-14 stream miles upstream of
the study site. An area of failed septic systems in the same
upstream vicinity also contributes to the pollutant loading of the
river.
Although water quality is degraded somewhat in the immedi-
ate vicinity of these municipal pollutant sources, as the river flows
southward towards Long Island Sound, it recovers to support a
valuable recreational trout fishery. There are no industrial point
source discharges of metals upstream of the study area.
Within the study area itself, the Gilbert and Bennett Manufac-
turing Company discharges treated process water to the Norwalk
River at a point below Factory Pond in Georgetown, Connecticut.
Gilbert and Bennett cleans, draws, and coats metal wire. Waste-
water is primarily generated during the wire cleaning process. The
wastewater treatment system of the facility consists of pH
neutralization and equalization followed by precipitation and
clarification of the effluent before discharge to the river. The
treated wastewater is discharged intermittently to the river.
The Connecticut Department of Environmental Protection un-
dertook the study of the Norwalk River site because the Gilbert
and Bennett metal loadings were calculated to result in excursions
of national water quality criteria for lead and zinc under design
flow conditions. In order to evaluate the effect of site water on the
toxicity of lead and zinc, EPA and the State used the indicator
species (water-effect ratio) protocol.
By testing a sensitive invertebrate and a fish in both site and
reconstituted laboratory dilution water, the water-effect ratio pro-
cedure accounts for differences in bioavailability and effective
toxicity of a pollutant in the two waters. The procedure responds
to the summation of all synergistic and antagonistic effects of site
water quality characteristics (including pH, hardness, particulate
matter, dissolved organic matter, and other contaminants). The
procedure does not, however, elucidate factors causing the dif-
ference in toxicity.
A water-effect ratio is the ratio of a species LC50 in site water
versus its LC50 in laboratory dilution water. The 1983 Water
Quality Standards Handbook recommends that two relatively sen-
sitive indicator species be tested, and that the geometric mean of
the two results be used. The site-specific criterion would be calcu-
lated as the product of the national criterion and the water-effect
ratio.
Introduction
13
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Analyses
Conducted
The results from the testing of metals toxicity in laboratory
and site water forms the primary basis for the site-specific
criteria. To provide additional information, the monitoring
of ambient water chemistry, surveying of macroinvertebrates, and
testing of whole effluent toxicity were also performed.
Chemical and Ecological Monitoring
Concentrations of several metals were measured in composite
samples taken at each of four ambient stations and in grab
samples of the Gilbert and Bennett main effluent. One ambient
station was in the upstream control zone, two in the impact zone,
and one in the recovery zone.
Benthic populations were sampled at five locations to assess
the impact of the discharge on the stream community. One refer-
ence station was located in the upstream control zone. Three
stations were in the Bennett and Gilbert impact zone, and one was
in the recovery zone. Physical substrate, stream velocity, and
water depth were similar at each location. Four Surber samples
were collected at each of the five locations. Organisms were sorted
in the field, preserved in 70 percent ethanol, and returned to the
laboratory for identification and enumeration.
Toxicity Testing
Norwalk River water was withdrawn from a station upstream of
Gilbert and Bennett and transported (along with the effluent
samples) back to the laboratory. Toxicity tests were conducted in
the sampled river water and in reconstituted water, with differing
amounts of either lead or zinc added, in order to determine the
LC50. Whole toxicity testing of one of the Gilbert and Bennett
effluents was also performed, using both upstream Norwalk River
water and reconstituted water for dilution.
Because the lead and zinc toxicity tests were run using
upstream water, they do not indicate the effects of synergism, an-
tagonism, or toxicant additivity with constituents of the Gilbert
and Bennett effluent. Although this guidance recommends use of
downstream water for at least one sample, the case study predates
the guidance and does not follow this recommendation.
Ninety-six hour acute toxicity tests (static with measured con-
centrations of toxicant) were conducted with laboratory reared
rainbow trout (Oncorhynchus mykiss, formerly Salmo gairdneri) and
48-hour acute toxicity tests (static with measured concentrations)
were conducted with laboratory-reared Daphnia magna.
14
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Water Chemistry and Ecological Quality
Mean instream concentrations of lead, zinc, and cadmium were
higher in the impact and recovery zones than in the control zone.
Levels of cadmium and copper appeared to exceed national acute
criteria at all sampling locations both upstream and downstream
of Gilbert and Bennett.
It should be noted that the metals data were generated using
the sample handling and analytical protocols of the early 1980s,
rather than more recent protocols emphasizing ultra-clean tech-
niques. While the link between chemical quality and ecological
quality is of great interest, it is not clear that these ambient metals
data are sufficiently reliable to be used in such comparisons. If the
ambient metals concentrations were reliably known, such infor-
mation would be most useful for comparing concentrations in the
control, impact, and recovery zones with the criteria derived from
the water-effect ratio.
At the upstream reference location, 889 organisms from 44 taxa
were collected. Most of the species collected could be cL^sified as
sensitive or facultative with respect to pollution tolerance. Species
diversity was high (Shannon Diversity index of 3.4), indicating ac-
ceptable water quality and aquatic habitat.
At the three near downstream locations, the number of or-
ganisms, number of taxa, and diversity were reduced significantly.
At some of the impact zone stations, the number of organisms was
less than one-fourth, and number of taxa one-third that of the ref-
erence site. The Shannon index registered as low as 1.0.
At the recovery zone station (500 m downstream from the dis-
charge), a larger number of organisms were found than at any of
the other stations, including the upstream reference site. Diversity
and numbers of taxa, however, remained at levels more charac-
teristic of the impacted stations.
The ecological assessment demonstrated that the ecology of
the Norwalk River was impaired, and strongly suggested that
some type of pollutant release from Gilbert and Bennett was in-
volved in the impairment. However, as noted in the Technical
Support Document [1], ecological assessments cannot identify the
causative agents, and generally do not predict the ecological
quality as a function of chemical-specific concentrations. Conse-
quently, these ecological data do not indicate the appropriateness
of particular values for the water-effect ratio.
Findings
Toxicity Testing with Lead and Zinc
Table A-l shows the results of static toxicity tests with daphnids
and rainbow trout exposed to lead and zinc, in river water and in
laboratory water.
15
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Table A-1.—Toxicity of lead and zinc in Norwalk River and in laboratory water.
DAPHNIA MAGNA
Lead
River water
Lab water
Zinc
River water
Lab water
48-hr LC50, nfl/L
(95% confid Mm)
1300(950-1900)
320 (290-360)
900(740-1100)
400 (380-480)
Water-Effect
Ratio
4.1
2.3
RAINBOW TROUT
96-hr LCSO, ng/L
(95% confid llm)
9600(7500-12000)
2600(1900-3600)
1500(1200-1800)
1000(850-1200)
Water-Effect
Ratio
3.7
1.5
The toxicity of both lead and zinc was lower in Norwalk River
water than in laboratory water. For both metals, the more sensitive
species, Daphnia magna, had the higher water-effect ratio. This is
consistent with general tendencies observed in other studies [10].
Whole Effluent Toxicity for One Effluent
In unspiked, static toxicity tests in which rainbow trout were
exposed to one of the Gilbert and Bennett effluents, the effluent
was rendered nontoxic by relatively little dilution of the effluent.
As effluent constituted the bulk (60-68 percent) of the test water at
dilutions toxic to half the organisms, it was not surprising that the
whole effluent toxicity tests could not discern differences between
toxicity in laboratory and river water. That is, toxicity was much
reduced before enough laboratory or river water could added to
the effluent to discern differences between the added water.
The monitored effluent was not sufficiently toxic to daphnids
to allow calculation of the dilution lethal to half the organisms.
However, to eliminate toxicity to all the tested individuals, sig-
nificantly more dilution was required with laboratory water than
with river water. This suggests that this effluent may be less toxic
in Norwalk River water than in laboratory water.
If the lead and zinc concentrations had been measured during
the whole effluent toxicity tests, it would be possible to compare
effect concentrations with the lead and zinc LC50 values shown in
the previous section. In making such comparisons, however, it
must be recognized that the cause of toxicity of the effluent is not
known.
Finally, it should be noted that the tested effluent is only one of
Gilbert and Bennett's releases to the Norwalk River. There is no
disparity between the observed significant instream impacts and
the relatively low toxicity of the one monitored effluent. The
ecological assessment suggested that the unmonitored release was
more toxic than the monitored one.
16
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Calculation of the Site-Specific Criteria
The water-effect ratios for lead and zinc differed relatively little
between species. If the overall water-effect ratio for each metal
were calculated from the geometric mean of the species
water-effect ratios, then the water-effect ratio for lead would be
3.9, and that for zinc would be 1.9.
It is assumed that the water-effect ratio would apply to both
the acute and the chronic criteria. As the national criteria for lead
and zinc are hardness dependent, for purposes of determining the
value of the site-specific criteria during the survey period, it is ap-
propriate to calculate the national criteria at the hardness of the
laboratory reconstituted water, if different from the site water. The
site-specific acute and chronic criteria for each metal would equal
the national (or state-wide) criteria multiplied by the water-effect
ratio for each metal.
Because all of the above toxicity tests and water-effect ratios
were based on total recoverable metal, the resulting site-specific
criteria would also be expressed as total recoverable metal.
When the site-specific criteria were used to calculate effluent
limits, it was found that large reductions in current metals load-
ings would be required. This result is not surprising, considering
the ecological effects observed downstream.
17
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APPENDIX B
Derivation of Effluent Limits
from Ambient Metals Criteria
-------
The determination of the waste loads and effluent limits that
allow attainment of water quality criteria is described in
other EPA guidance [1, 11, 12]. The Waste Load Allocation
Guidance Manuals [11, 12] are particularly suited to predicting
far-field dissolved and total recoverable concentrations, such as
would be necessary for evaluating watersheds with multiple dis-
chargers. The Technical Support Document (TSD) [1] has
additional guidance on the evaluation of mixing zones and the
derivation of permit limits. Nevertheless, some additional discus-
sion is provided below. This guidance supersedes the second
paragraph of Section 5.7.3 "Metals" of the TSD [I].
Expressing state-wide or site-specific criteria as total
recoverable metal has the advantage of providing a simple
basis for calculating effluent limits. All of the effluent total
recoverable metal would contribute to the ambient total
recoverable concentration.
f criteria are expressed as dissolved metal, then it is necessary to
establish what fraction of the effluent total recoverable metal
contributes to the ambient dissolved metal.
Three alternatives may be considered for relating the ambient
dissolved criterion for a specific metal to the effluent total
recoverable limits: (a) directly measure dissolved and total
recoverable metal in the receiving water, (b) assign environmental-
ly conservative default values for the assumed ratio between
dissolved and total recoverable metal, and (c) predict the percent-
age dissolved metal from a geochemical model such as MINTEQ.
Regardless of which alternative is used, it must be recognized
that the goal is to set the effluent limit at a value such that the am-
bient water quality standard will be attained. In addition,
compliance with regulatory requirements for technology-based
limits, antidegradation, and antibacksliding is necessary.
Using Site-Specific Measurements
The concept is to measure the dissolved-total ratio for the
particular metal in the receiving water affected by the discharger.
Because the chemical properties of an effluent (particularly an
industrial effluent) may be much different than the chemical
properties of the receiving water below the discharger, it is
appropriate to measure the ratio in the receiving water rather than
in the effluent. As an approximation, it may be assumed that the
measured dissolved-total metal ratio in the affected waters reflects
the fraction of effluent total recoverable metal that remains or
becomes dissolved in ambient water.
Total
Recoverable
Metal Criteria
Dissolved
Metal Criteria
21
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Samples on which measurements are made should be repre-
sentative of the bulk of the receiving water. It is recommended that
sampling be performed over a period of time, with samples repre-
senting the usual range of effluent and ambient quality, while
emphasizing the season corresponding to the critical water quality
conditions. Because the control strategy assumes that the dis-
solved concentration is related to the total recoverable
concentration, it would be appropriate to verify that the dissolved
and total recoverable concentrations are in fact correlated.
In freshwaters, an alternative approach to downstream sam-
pling is to sample the effluent and the upstream waters and mix
samples at an appropriate dilution. The dilution and the seasons
for sampling should be related to the critical conditions, although
it may be appropriate to reduce the dilution if necessary to detect
and quantify the metal.
The most important constraint on the feasibility of carrying
out site-specific measurements is the capability of analytical
laboratories to detect and accurately quantify both the dissolved
and total recoverable metal. Graphite furnace (flameless) atomic
absorption AA techniques are usually needed. Furthermore, great
care is needed to prevent external contamination of samples. The
EPA- and USGS-recommended sample handling methods, com-
monly used, may produce inaccurate results when judged against
newer techniques that emphasize highly purified reagents, Teflon
and polyethylene labware, and clean laboratory environments [7,
13].
The high degree of imprecision of metals measurements tends
to result in overstatement of the true variability of the dissolved-
total ratio. As a result, unless a mean or median observed ratio is
used, it may be necessary to compensate for the effect of measure-
ment imprecision (by subtracting out the measurement
imprecision variance).
In order to provide some sense of the general magnitude of
typically observed dissolved-total metals ratios, data from several
sources have been compiled in Table B-l. The ambient data under-
lying the tabulated values are considered to be reasonably reliable.
Using Environmentally Conservative Default Values
This option is best applied as the first tier of a tiered approach,
where the second tier involves site-specific measurements. In this
type of framework, the default (first-tier) percentage dissolved
might be set at a reasonable worst-case value.
One possible worst-case assumption is that 100 percent of the
effluent total recoverable metal will become dissolved in the
receiving water. Such an assumption may be particularly ap-
propriate for a metal, such as mercury, for which there are
substantial uncertainties regarding long-term processes convert-
22
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Table B-1.—Observed average fractions of dissolved metals in ambient
waters.
METAL
Aluminum
Cadmium
Copper
Lead
Nickel
Silver
Zinc
FRESH WATER ~
EAST COAST [7]
0.40
0.62
0.10
0.20
SALT WATER
NY-NJ HARBOR AREA [13]
NEAR SURFACE
0.81b
0.50°
0.08C
0.73°
0.11C
0.60d
NEAR BOTTOM
0.56°
0.23d
0.03d
0.41d
0.08C
0.29d
FRESH
AND SALT
WATER
STORET"
0.078
0.4f
0.6
"For STORE! data, means were estimated from cumulative distributions of concentrations, 1984-
1990, in ambient streams, rivers, canals, lakes, reservoirs, and estuaries, for samples in which
both dissolved and total recoverable metal were analyzed. The fraction dissolved for STORET
data for cadmium, lead, silver, and zinc are not tabulated because most of these data are
believed to be seriously compromised by external contamination of samples.
"Dissolved and total recoverable concentrations well correlated.
Dissolved and total recoverable concentrations somewhat correlated.
"Dissolved and total recoverable concentrations not correlated.
eFor any particular measurement of aluminum, the method of filtration may strongly affect the
result.
°For much of the STORET copper data, external contamination of samples is likely to somewhat
affect the absolute values of the measured concentrations, and may somewhat affect the dis-
solved-total ratio.
ing inorganic (including particulate) mercury into bioaccumula-
tive methyl mercury. Where the background metal concentration
is either negligible or entirely dissolved, and where the dissolved
criterion is less than the corresponding total recoverable criterion,
the assumption that all effluent total recoverable metal will be-
come dissolved yields more restrictive limits than simply
implementing a total recoverable criterion.
Other environmentally conservative default values may also
be developed based on available information. Such values may
differ in different parts of the country, due to variations in water
quality characteristics.
Using a Geochemical Model
The equilibrium metal speciation model MINTEQ may offer
assistance in understanding or predicting the fraction dissolved
[14]. Using MINTEQ without obtaining site-specific input data
may not be feasible, however. As the effort in obtaining the
appropriate input data would likely be equivalent to simply
measuring the site-specific fraction dissolved, the model may be
more useful for providing insight into the controlling factors and
predicting the effects of different environmental conditions.
MINTEQ may be particularly useful for predicting whether the
dissolved concentration will respond to reductions in the total
concentration or whether it is controlled by factors such as
solubility.
' ;-ivrarv (FL-l?J»
J;c!iSL-?i.&t;i?je!/a'd( 12th
II 63694-3590
23
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