<pubnumber>822R01001</pubnumber>
<title>2001 Update of Ambient Water Quality Criteria for Cadmium</title>
<pages>276</pages>
<pubyear>2001</pubyear>
<provider>NEPIS</provider>
<access>online</access>
<operator>dwu</operator>
<scandate>04/26/01</scandate>
<origin>hardcopy</origin>
<type>single page tiff</type>
<keyword>cadmium environ toxicol toxicity acute effects bull freshwater metals water aquat chronic species fish hardness mar metal accumulation uptake trout</keyword>
United States
Environmental Protection
Agency
Office of Water
(4304)
EPA-822-R-O1-OO1
April 2001
2001 Update of
Ambient Water
Quality Criteria for
Cadmium
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EPA-822-R-01-001
April 2001
2001 UPDATE OF AMBIENT WATER QUALITY CRITERIA FOR
CADMIUM
(CAS Registry Number 7440-43-9)
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, D.C.
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NOTICES
This document has been reviewed by the Health and Ecological Criteria Division, Office of Science and
Technology, U.S. Environmental Protection Agency, and is approved for publication.
Mention of trade names or commercial products does not constitute endorsement or recommendation
for use.
This document is available to the public through the National Technical Information Service (NTIS),
5285 Port Royal Road, Springfield, VA 22161.
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ACKNOWLEDGMENTS
Document Update: 1984
John G. Eaton
(freshwater author)
Environmental Research Laboratory
Duluth, Minnesota
Charles E. Stephan
(document coordinator)
Environmental Research Laboratory
Duluth, Minnesota
Statistical Support:
Clerical Support:
John W. Rogers
Terry L. Highland
John H. Gentile
(saltwater author)
Environmental Research Laboratory
Narragansett, Rhode Island
David J. Hansen
(saltwater coordinator)
Environmental Research Laboratory
Narragansett, Rhode Island
Document Update: 2001
Gregory J. Smith
(freshwater contributor)
Great Lakes Environmental Center
Columbus, Ohio
Cindy Roberts
(document coordinator)
U.S. EPA
Health and Ecological Effects Criteria Division
Washington, D.C.
Statistical Support: Dan Tholen, Great Lakes Environmental Center, Traverse City, Michigan
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CONTENTS
Notices ii
Acknowledgments Hi
Contents iv
Tables v
Figures vi
Introduction 1
Acute Toxicity to Freshwater Animals 3
Acute Toxicity to Saltwater Animals 9
Chronic Toxicity to Freshwater Animals 12
Chronic Toxicity to Saltwater Animals 16
Toxicity to Aquatic Plants '. 18
Bioaccumulation 18
Other Data 20
Unused Data 22
Summary 30
National Criteria 31
References .- '. . 156
iv
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TABLES
la. Acute Toxicity of Cadmium to Freshwater Animals 40
Ib. Acute Toxicity of Cadmium to Saltwater Animals 56
Ic. Results of Covariance Analysis of Freshwater Acute Toxicity Versus Hardness 65
Id. List of Studies Used to Estimate Acute Cadmium Hardness Slope 66
2a. Chronic Toxicity of Cadmium to Freshwater Animals 74
2b. Chronic Toxicity of Cadmium to Saltwater Animals 77
2c. Results of Covariance Analysis of Freshwater Chronic Toxicity Versus Hardness 78
2d. List of Studies Used to Estimate Chronic Cadmium Hardness Slope 78
2e. Cadmium Acute-Chronic Ratios . 79
3a. Ranked Freshwater Genus Mean Acute Values with Species Mean Acute-Chronic Ratios . 80
3b. Ranked Saltwater Genus Mean Acute Values with Species Mean Acute-Chronic Ratios . . 86
3c. Ranked Freshwater Genus Mean Chronic Values 92
3d. Freshwater and Saltwater Cadmium Criteria Values 94
4a. Toxicity of Cadmium to Freshwater Plants 97
4b. Toxicity of Cadmium to Saltwater Plants 100
5a. Bioaccumulation of Cadmium by Freshwater Organisms 101
5b. Bioaccumulation of Cadmium by Saltwater Organisms . 107
6a. Other Data on Effects of Cadmium on Freshwater Organisms 109
6b. Other Data on Effects of Cadmium on Saltwater Organisms 143
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FIGURES
1. Comparison of All Table 1 Freshwater Acute Toxicity Test EC50s and LC50s with the
Hardness Slope Derived CMC (2001 CMC: solid line; 1984 CMC: dashed line) 33
2. Ranked Summary of Cadmium GMAVs (Freshwater) 34
3. Ranked Summary of Cadmium GMAVs (Saltwater) 35
4. Comparison of All Table 2 Freshwater Chronic Values with the Hardness Slope Derived CCC
(2001 CMC: solid line; 1984 CMC: dashed line) 36
5. Chronic Toxicity of Cadmium to Aquatic Animals 37
6. Comparison of Freshwater Plant Toxicity Values (Table 4) and Freshwater CMC and
CCC Values 38
7. Comparison of Saltwater Plant Toxicity Values (Table 4) and Saltwater CMC and
CCC Values 39
VI
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INTRODUCTION1
This update document provides guidance to States and Tribes authorized to establish water
quality standards under the Clean Water Act (CWA) to protect aquatic life from acute and chronic
effects of cadmium. Under the CWA, States and Tribes are to establish water quality criteria to protect
designated uses. While this document constitutes.U.S. EPA's scientific recommendations regarding
ambient concentrations of cadmium, this document does not substitute for the CWA or U.S. EPA's
regulations; nor is it a regulation itself. Thus, it cannot impose legally binding requirements on U.S.
EPA, States, Tribes, or the regulated community, and might not apply to a particular situation based
upon the circumstances. State and Tribal decision-makers retain the discretion to adopt approaches on a
case-by-case basis that differ from this guidance when appropriate. U.S. EPA may change this guidance
in the future.
Cadmium is a relatively rare element that is a minor nutrient for plants at low concentrations
(Lane and Morel 2000; Lee et al. 1995; Price and Morel 1990), but is toxic to aquatic life at
concentrations only slightly higher. It occurs mainly as a component of minerals in the earth's crust at
an average concentration of 0.18 ppm (Babich and Stotzky 1978). Cadmium levels in soils usually
range from approximately 0.01 to 1.8 ppm (Lagerwerff and Specht 1970). In natural freshwaters,
cadmium sometimes occurs at concentrations of less than 0.1 /tg/L, but in environments impacted by
man, concentrations can be several micrograms per liter or greater (Abbasi and Soni 1986; Allen 1994;
Annune et al. 1994; Flick et al. 1971; Friberg et al. 1971; Henriksen and Wright 1978; Nilsson 1970;
Spry and Wiener 1991). Cadmium can enter the environment from various anthropogenic sources,
such as by-products from zinc refining, coal combustion, mine wastes, electroplating processes, iron
and steel production, pigments, fertilizers and pesticides (Hutton 1983; Pickering and Gast 1972).
The impact of cadmium on aquatic organisms depends on a variety of possible chemical forms
of cadmium (Callahan et al. 1979), which can have different toxicities and bioconcentration factors. In
most well oxygenated freshwaters that are low hi total organic carbon, free divalent cadmium will be
the predominant form. Precipitation by carbonate or hydroxide and formation of soluble complexes by
chloride, sulfate, carbonate, and hydroxide should usually be of little importance. In saltwaters with
An understanding of the "Guidelines for Deriving Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and Their Uses" (Stephan et al. 1985), hereafter referred to as the Guidelines, is
necessary in order to understand the following text, tables, and calculations.
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salinities from about 10 to 35 g/kg, cadmium chloride complexes predominate. In both fresh and
saltwaters, participate matter and dissolved organic material may bind a substantial portion of the
cadmium, and under these conditions cadmium may not be bioavailable due to this binding (Callahan et
al. 1979; Kramer et al. 1997).
Because of the variety of forms of cadmium (Callahan et al. 1979) and lack of definitive
information about their relative toxicities, no available analytical measurement is known to be ideal for
expressing aquatic life criteria for cadmium. Previous aquatic life criteria for cadmium (U.S. EPA
1980) were expressed in terms of total recoverable cadmium (U.S. EPA 1983a), but this measurement
is probably too rigorous in some situations. U.S. EPA (1985) has also expressed cadmium criteria as
acid-soluble cadmium in the past, but now recommends use of dissolved metal concentrations
(operationally defined as the metal in solution that passes through a 0.45 ^m membrane filter) to set and
measure compliance with water quality standards (Prothro 1993; U.S. EPA 1993, 1994a).
The criteria presented herein supersede previous aquatic life water quality criteria for cadmium
(U.S. EPA 1999a) because these new criteria were derived based on the most recent science.
Whenever appropriate, a national criterion may be replaced by a site-specific criterion (U.S. EPA
1994a), which may include not only site-specific criterion concentrations (U.S. EPA 1994b), but also
site-specific durations of averaging periods and site-specific frequencies of allowed exceedences (U.S.
EPA 1991). All concentrations are expressed as cadmium, not as the chemical tested. The latest
literature search for information for this document was conducted in June 1999; some newer
information was also used.
Because the revisions being considered build from principles set forth in the 1985 Guidelines
(Stephen et al. 1985), it is useful to have some understanding of how those Guidelines are ordinarily
applied: (1) Acute toxicity test data must be available for species from a nunimum of eight diverse
taxonomic groups. The diversity of tested species is intended to assure protection of various
components of an aquatic ecosystem. (2) The Final Acute Value (FAV) is derived by extrapolation or
interpolation to a hypothetical genus more sensitive than 95 percent of all tested genera. The FAV,
which represents an LC50 or EC50, is divided by two in order to obtain an acute criterion protective of
nearly all individuals in such a genus. (3) Chronic toxicity test data (longer-term survival, growth, or
reproduction) must be available for at least three taxa. Most often the chronic criterion is set by
determining an appropriate acute-chronic ratio (the ratio of acutely toxic concentrations to the
chronically toxic concentrations) and applying that ratio to the acute value of the hypothetical genus
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more sensitive than 95 percent of all tested genera. If sufficient data are available to meet the eight
diverse taxonomic group minimum, then the chronic value is derived using the same procedure as used
for the FAV derivation. (4) When necessary, the acute and/or chronic criterion may be lowered to
protect recreationally or commercially important species. (5) When evaluating tune-variable ambient
concentrations generally, 1-hour average concentration are considered to be appropriate for comparison
with the acute criterion, and 4-day averages with the chronic criterion. (6) The allowable frequency for
exceeding a criterion is set at once every three years, on the average.
ACUTE TOXICITY TO FRESHWATER ANIMALS
Acceptable data on the acute effects of cadmium in freshwater are available for 39 species of
invertebrates, 24 species of fish, one salamander species, and one frog species (Table la). These 65
species satisfy the eight different family requirements specified in the Guidelines. A tendency for
increased tolerance to toxicity with increasing size or age has been reported (Table la) in the snail,
Physa gyrina (Wier and Walter 1976), the coho salmon (Chapman 1975), and the common carp (Suresh
et al. 1993a). No such effect was observed with increasing age (Table la) in the cladocerah, Daphnia
magna (Stuhlbacher et al. 1993), the rainbow trout (Chapman 1975, 1978), or in the striped bass
(Hughes 1973; Palawski et al. 1985). Data are unavailable for a sufficient number of species and life
stages to allow general adjustment of test results or criteria on the basis of size or life stage. Where
relationships were apparent between life-stage and sensitivity, only values for the most sensitive life-
stage were considered.
Water Quality Parameters Affecting Toxicity
Although many factors might affect the results of tests of the toxicity of cadmium to aquatic
organisms (Sprague 1985), water quality criteria can quantitatively take into account only factors for
which enough data are available to show that the factor similarly affects the results of tests with a
variety of species. Hardness is often thought of as having a major effect on the toxicity of cadmium,
although the observed effect may be due to one or more of a number of usually interrelated ions, such
as hydroxide, carbonate, calcium, and magnesium. Acute tests were conducted at three different levels
of water hardness with Daphnia magna (Chapman et al. Manuscript), demonstrating that daphnids were
at least five tunes more sensitive to cadmium in soft water than in hard water (Table la). Data in Table
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la also indicate that cadmium was more toxic to the tubificid worms Limnodrilus hoffmeisteri and
Tubifex tubifex, the mussel Vilosa vibex, Daphnia pulex, chinook salmon, goldfish, fathead minnow,
guppy, striped bass, green sunfish and bluegill in soft than in hard water. Carroll et al. (1979) found
that calcium, but not magnesium, reduced the acute toxicity of cadmium.
Other water quality characteristics could potentially influence the toxicity of cadmium to aquatic
species. Giesy et al. (1977) found that dissolved organics substantially reduced the toxicity of cadmium
to daphnids, but had little effect on its toxicity to fish. No consistent relationship between toxicity and
organic particle size was observed. Development of the "biotic ligand model" (BLM - formerly the
"gill model") hi recent years has attempted to better account for the bioavailability of metals to aquatic
life. The BLM, which quantifies the capacity of metals to bind to the gills of aquatic organisms, has
been proposed as a reliable method for estimating the bioavailable portion of dissolved metals hi the
water column based on site-specific water quality parameters such as alkalinity, pH and dissolved
organic carbon (McGeer et al. 2000; Meyer et al. 1999; Pagenkopf 1983; Paquin et al. 1999; U.S.
EPA 1999b, 2000). Future development of the BLM for cadmium may help better quantify the
bioavailable fraction of cadmium. Nonetheless, the model is in the preliminary development phase for
cadmium and it will likely not be available for a number of years still.
Hardness Correction
Currently, the primary quantitative correlation used to modify metal toxicity estimates is water
hardness (viz. the U.S. EPA 1995 water quality criteria for cadmium). Hardness (as calcium or
magnesium ions) almost certainly has some direct effect on cadmium toxicity (e.g., by influencing
membrane integrity). Calcium and magnesium ions compete with the metal for binding sites on the gill
(Carroll et al. 1979; Evans 1987; Morel and Bering 1993; Pagenkopf 1983). Hardness also serves as a
general surrogate for pH, alkalinity, and ionic strength, because waters of higher hardness usually have
higher pH, alkalinity, and ionic strength. Other parameters such as pH, alkalinity, dissolved organic
carbon, humic matter, ionic strength (anions and cations) and dissolved inorganic carbon also affect
metal speciation and bioavailability, and thus metal toxicity. The pH is also important hi determining
the metal complexation capacity of dissolved organic matter.
Hardness is used here as a surrogate for the ions which affect the results of toxicity tests on
cadmium. However, it should be emphasized that the hardness adjustment is not a precise measure, but
an estimation. The variability associated with different life stages, clones and test conditions of the
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studies used to determine the hardness slope all contribute to the uncertainty of the hardness correction.
In selected cases, only one life stage was used in the analysis (e.g., only adult fathead minnow data).
Thus, in spite of all its limitations, hardness is currently the best surrogate available for metal toxicity
adjustment.
To account for the apparent relationship of cadmium acute toxicity to hardness, an analysis of
covariance (Dixon and Brown 1979; Neter and Wasserman 1974) as noted hi the guidelines (Stephan et
al. 1985) was performed using the Statistical Analysis System (SAS Inc., Gary, NC) software program
to calculate the pooled slope for hardness using the natural logarithm of the acute value as the
dependent variable, species as the treatment or grouping variable, and the natural logarithm of hardness
as the covariate or independent variable. The pooled slope is a regression slope from a pooled data set,
where every variable is adjusted relative to its mean. The species are adjusted separately, then pooled
for a single conventional least squares regression analysis. The slope of the regression line is the best
estimate of the all-species relationship between toxicity and hardness. With analysis of covariance,
different species will be weighted relative to the number of data points they have. In this case, the D.
magna and the fathead minnow each have 28 data points out of the total of 97, and the next most
frequent species has just eight data points.
This analysis of covariance model was fit to the data in Table la for the 12 species for which
definitive acute values (less than or greater than values were not used) are available over a range of
hardness such that the highest hardness is at least three tunes the lowest, and the highest is also at least
100 mg/L higher than the lowest (other species in Table la either did not meet these criteria or did not
show any hardness-toxicity trend due to differences in exposure methods, species age, etc.). For D.
magna, only acute toxicity tests that were initiated with less than 24-hr old neonates were used to
estimate the hardness slope. For the fathead minnow, only tests conducted with adults were used (not
those conducted with the more sensitive fry life stage). A list of the species and acute toxicity-hardness
values used to estimate the acute hardness slope is provided in Table Id. The slopes for all 12 species
ranged from 0.1086 to 2.031, and the pooled slope for these 12 species was 1.174 (see Table Ic). An
F-test was used to test whether a model with separate species slopes for each species gives significantly
better fit to the data than the model with parallel slopes. This test showed that the separate slopes
model is not significantly better, and therefore the slopes are not significantly different than the overall
pooled slope (P=0.27). The slopes and confidence intervals associated with the 12 species indicated
that D. magna (all available data) had a very flat slope and a large confidence interval (and large
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standard error). If only the D. magna data from Chapman et al. (Manuscript) were used, the resultant
D. magna slope was 1.182, with smaller confidence intervals than for the all D. magna. slope.
Likewise, when only the adult fathead minnow data were used (not the fry data), the resultant fathead
minnow slope was 1.221 and smaller confidence intervals were present. If this reduced data set is used
(all species but using only data from Chapman et al. (Manuscript) for D. magna and only adult fathead
minnow data), the pooled slope for these species was 1.0166 (see Table Ic). The test for equality of
the 12 slopes using the reduced data set (all species but only Chapman D. magna and adult fathead
minnow data) produced P=0.69. Under analysis of covariance, it therefore is reasonable to assume
that the slopes for these 12 species are the same, and that the overall slope is a reasonable estimate of
the average relationship between hardness and toxicity. Either P value indicated that it was reasonable
to assume that the slopes were the same, however, the second model was considered the better model
and was therefore selected. The pooled slope of 1.0166 is close to the slope of 1.0 that is expected on
the basis that cadmium, calcium, magnesium, and carbonate all have a charge of two (Meyer 1999). A
plot of the acute effect level (EC50 or LC50) versus total hardness is provided ha Figure 1.
The possible relationship of cadmium acute toxicity to water quality parameters other than
hardness were also considered. Both hardness and/or alkalinity were investigated by subjecting any
acute toxicity data hi Table la having both hardness and alkalinity values available to a multiple
stepwise regression analysis using the SAS (Gary, NC) software program. The analysis was run using
the natural logarithm of the acute value as the dependent variable, species as the treatment or grouping
variable, and the natural logarithm of hardness and alkalinity as the covariates or independent variables.
As with the analysis of covariance evaluation discussed above, the only data used hi Table la (seven
species) were those for which definitive acute values are available over a range of both hardness and
alkalinity such that the highest hardness (and alkalinity) is at least three times the lowest, and the
highest is also at least 100 mg/L higher than the lowest. The results obtained indicate that either
variable works well alone in the regression model (R2 value for each was 0.688), but the other variable
cannot increase the strength of the model once the first variable is included (when both were used the R2
value only increased to 0.689). This lack of model improvement is due to the very strong correlation
between hardness and alkalinity (effect of colinearity), thus these two independent variables should not
be used together hi the same regression model. Based on these results and the availability of data for
water quality parameters other than hardness, the best approach at this tune is to use only hardness
(analysis of covariance discussed above) as a surrogate for the ions which affect the results of toxicity
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tests on cadmium.
Conversion Factors
Although past water quality criteria for cadmium (and other metals) have been established upon
the loosely defined term of "acid soluble metals," U.S. EPA made the decision to allow the expression
of metal criteria on the basis of dissolved metal (U.S. EPA 1994a), operationally defined as that metal
that passes through a 0.45 micron filter. Because most of the data in existing databases are from tests
that were either nominal concentrations, or provided only total cadmium measurements, some
procedure was required to estimate then- dissolved equivalents. The approach taken by U.S. EPA
involves the use of conversion factors (CF), that when applied to the total metal concentration, gives a
dissolved metal concentration. Thus, the CF corresponds to the percent of the total recoverable metal
that is dissolved. These CFs were determined by conducting a number of "simulation tests" using
solutions simulating those used hi the toxicity tests that were most important hi the derivation of aquatic
life criteria for each metal (static, flow-through, fed, and unfed conditions that typified standard acute
and chronic toxicity tests from which criteria are derived). The intent was to mimic the way criteria
would have been derived if dissolved metal had been measured hi each of the toxicity tests (Lussier et
al. 1995; Stephan 1995; Univ. of Wisconsin-Superior 1995). For certain metals like cadmium, these
CFs are hardness dependent.
The appropriate CFs were used only when determining the final cadmium criteria values, and
are hardness dependent hi freshwater. Acute freshwater total cadmium concentrations were converted
to dissolved concentrations using the factor of 0.973 at a total hardness level of 50 mg/L as CaCO3,
0.944 at a total hardness level of 100 mg/L as CaCO3, and 0.915 at a total hardness level of 200 mg/L
as CaCO3. The equation for the acute freshwater conversion factor is CF = 1.136672 - [(hi hardness)
(0.041838)] where the (hi hardness) is the natural logarithm of the hardness (Stephen 1995). Acute
saltwater total cadmium values were converted to dissolved using the factor of 0.994.
Criteria Development
The pooled slope of 1.0166 was used to adjust the freshwater acute values hi Table la to
hardness = 50 mg/L, except where it was not possible because no hardness was reported. Species
Mean Acute Values (SMAV) were calculated as geometric means of the adjusted acute values (only the
underlined EC50/LC50 species values were used to calculate the respective SMAV). As stated hi the
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Guidelines (Stephen et al. 1985), flow-through measured study data are normally given preference over
non-flow-through data for a particular species. In certain cases flow-through measured results were
available, yet preference was given to the sensitive life stage for certain species in calculating SMAVs.
In addition, all underlined Table la data for D. magna and fathead minnow fry were used to calculate
the respective SMAVs (D. magna tests initiated with > 24-hr old neonates were not used to calculate
the SMAV). Only data from Chapman (1975) were used for coho salmon to avoid using test results
from studies in which the life stage tested is known to be less sensitive, or in which the life stage tested
is unreported and the higher LC50s may be due primarily to the use of less sensitive life stages. The
data for Palawski et al. (1985) were used for striped bass because they were considered better data than
those given in U.S. EPA (1985), although the data from Hughes (1973) support the newer data. Only
brook trout data reported by Carroll et al. (1979), and not by Holcombe et al. (1983) were used in the
calculation of the brook trout Final Acute Value because the reported bull trout data (Stratus Consulting
1999) in the same genus support the Carroll et al. (1979) results. Drummond and Benoit (Manuscript)
reported that stress greatly affected the sensitivity of brook trout to cadmium.
The SMAV for freshwater invertebrates ranged from 13.41 jig/L total cadmium for the
cladoceran, D. magna to 96,880 jig/L total cadmium for the midge, Chironomus riparius. Of the fish
species tested, the brown trout, Salmo tnttta, had the lowest SMAV of 1.613 /*g/L total cadmium, and
the tilapia, Oreochromis mossambica, recorded the highest fish SMAV of 10,663 /tg/L total cadmium.
As indicated by the data, both invertebrate and fish species display a wide range of sensitivities to
cadmium.
Fish species represent eight of the nine most sensitive species to cadmium (Table 3a).
Salmonids (Salmo trutta, Salvelinus confluentus, Salvelinus fontinalis, Oncorhynchus Idsutch,
Oncorhynchus mykiss and Oncorhynchus tshawytscnd) are six of the seven most sensitive species listed
in Table la, and thus are more acutely sensitive to cadmium than any other freshwater animal species
thus far tested (Carroll et al. 1979; Chapman 1975, 1978, 1982; Cusimano et al. 1986; Davies et al.
1993; Finlayson and Verrue 1982; Phipps and Holcombe 1985; Spehar and Carlson 1984a,b; Stratus
Consulting 1999). The cladoceran, D. magna, is the eighth most sensitive species to cadmium, and
thus the most acutely sensitive invertebrate species tested thus far.
Genus Mean Acute Values (GMAV) at a hardness of 50 mg/L were then calculated (Table 3a)
as geometric means of the available freshwater Species Mean Acute Values and ranked. Of the 55
genera for which acute values are available, the most sensitive genus, Salmo, is over 60,062 times more
8
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sensitive than the most resistant, Chironomus. The first through fourth most sensitive genera (a total n
of 55) were used in the computation of the final acute value. The sensitivity of these four most
sensitive genera are within a factor of 2.4, and all are fish. Of the ten most sensitive genera, six are
fish, two are mussels, and two are cladocerans (Figure 2; Table 3a). Hardness-adjusted acute values
are available for more than one species in nine genera, and the range of SMAVs within each genus is
less than a factor of 4.0 for eight of the nine genera. The ninth genus, Ptychocheilus, has two SMAVs
that differ by a factor of 98.5, possibly due to differences hi the test conditions between species.
The freshwater Final Acute Value (FAV) for total cadmium at a hardness of 50 mg/L was
calculated to be 2.763 /^g/L total cadmium (Table 3d) from the Genus Mean Acute Values in Table 3a
using the procedure described in the Guidelines. The Species Mean Acute Values for the rainbow
trout, brook trout, bull trout and brown trout are lower than the FAV of 2.763 ^g/L total cadmium, but
the acute value for the brook trout and brown trout are from static tests, whereas flow-through
measured tests have been conducted with the remaining two salmonid species. The freshwater Final
Acute Value for total cadmium at a hardness of 50 mg/L was lowered to 2.108 /^g/L to protect the
commercially important rainbow trout (Table 3d). This value is above the SMAV of 1.613 jig/L for
the brown trout and < 1.791 /tig/L for brook trout, but below all other SMAVs listed in Table 3a
(Figure 2). The resultant freshwater Criterion Maximum Concentration (CMC) at a hardness of 50
mg/L for total cadmium (in ^g/L) = eU-Oi<56tin(hardness)]-3.924) If ^ CMC baged on total cadmium
values is converted to dissolved cadmium using the 0.973 factor at a hardness of 50 mg/L determined
by U.S. EPA (Stephan 1995; Univ. of Wisconsin-Superior 1995), the freshwater CMC for dissolved
cadmium (in ^g/L) = 0.973 [^^^^^^-^\ T^ me L0 Mg/L CMC for dissolved
cadmium at a hardness of 50 mg/L is below all of the SMAVs presented hi Table 3a (Figure 2).
Conversion from total to dissolved was used because hardness relationships were established based
upon total cadmium concentrations as this minimized the number of conversions required. In a few
cases where only dissolved cadmium was reported hi freshwater (Table la), conversion to total used the
same appropriate factor.
ACUTE TOXICITY TO SALTWATER ANIMALS
Tests of the acute toxicity of cadmium to saltwater organisms have been conducted with 50
species of invertebrates and 11 species offish (Table Ib), representing the required eight different
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taxonomic families. A pattern of increased tolerance to toxicity with increasing size or age has been
reported (Table Ib) in the polychaete worm Capitella capitata (Reish and LeMay 1991; Reish et al.
1976), the blue mussel (Ahsanullah 1976; Martin et al. 1981; Nelson et al. 1988), the copepod
Eurytemora affinis (Gentile 1982; Sullivan et al. 1983), the amphipods Marinogammarus obtusatus
(Wright and Frain 1981) and Leptocheirus plumulosus (McGee et al. 1998), the pink shrimp Penaeus
duoranan (Nimmo et al. 1977b; Gripe 1994), the rivulus (Park et al. 1994; Lin and Dunson 1993), the
Atlantic silverside (Cardin 1982) and the striped mullet (Hilmy et al. 1985). No such effect was
observed with increasing age (Table Ib) in the polychaete worm Neanthes arenaceodentata (Reish and
LeMay 1991; Reish et al. 1976), the mysid Americamysis bahia, formerly Mysidopsis bdhia (De Lisle
and Roberts 1988), the grass shrimp Palaemonetes pugio (Khan et al. 1988; Burton and Fisher 1990),
and the mummichog Fundulus heteroclitus (Voyer 1975). Data are unavailable for a sufficient number
of species and life stages to allow general adjustment of test results or criteria on the basis of size or life
stage. Where relationships were apparent between life-stage and sensitivity, only values for the most
sensitive life-stage were considered.
Water Quality Parameters Affecting Toxicity
Frank and Robertson (1979) reported that the acute toxicity to juvenile blue crabs was related to
salinity. The 96-hr LC50s were 320, 4,700, and 11,600 Aig/L at salinities of 1, 15, and 35 g/kg,
respectively (Table Ib). Studies with A bahia by Gentile et al. (1982) and Nimmo et al. (1977a) also
support a relationship between salinity and the acute toxicity of cadmium. O'Hara (1973a) investigated
the effect of temperature and salinity on the toxicity of cadmium to the fiddler crab. The LCSOs at
20°C were 32,300, 46,600, and 37,000 y/g/L at salinities of 10, 20, and 30 g/kg, respectively.
Increasing the temperature from 20 to 30°C lowered the LC50 at all salinities tested. Toudal and
Riisgard (1987) reported that increasing the temperature from 13 to 21 °C at a salinity of 20 g/kg also
lowered the LC50 value of cadmium to the copepod, Acartia tonsa.
Saltwater fish species were generally more resistant to cadmium than freshwater fish species
with SMAVs ranging from 75.0 //g/L for the striped bass (at a salinity of 1 g/kg) to 50,000 y.g/L for
the sheepshead minnow (Table 3b). In a study of the interaction of dissolved oxygen and salinity on the
acute toxicity of cadmium to the mummichog, Voyer (1975) found that 96-hr LCSOs at a salinity of 32
g/kg were about one-half what they were at 10 and 20 g/kg. Sensitivity of the mummichog to acute
cadmium poisoning was not influenced by reduction hi dissolved oxygen concentration to 4 mg/L. This
10
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increase in toxicity with increasing salinity conflicts with othef data reported in Tables Ib and 6b.
Since there was no consistent salinity-toxieity trend observed for the data, a salinity correction factor
was not attempted.
Criteria Development
Of the 54 saltwater genera for which acute values are available, the most sensitive,
Americamysis, is 3,270 times more sensitive than the most resistant, Monopylephorus (Table 3b). The
SMAVs for saltwater invertebrate species range from 41.29 ^g/L for a mysid to 135,000 ^g/L for an
.oligochaete worm (Tables Ib and 3b). The acute values for saltwater polychaetes range from 200 ^g/L
for C. capitata to 14,100 yWg/L for N. arenaceodentata (Reish and LeMay 1991). Saltwater molluscs
have Species Mean Acute Values from 227.9 //g/L for the Pacific oyster to 19,170 ^g/L for the mud
snail. Acute values are available for more than one species hi each of seven genera, and the range of
Species Mean Acute Values within each genus is no more than a factor of 3.6 for six of the seven
genera. The seventh genus, Crassostrea, has two SMAVs that differ by a factor of 16.7, possibly due
to different exposure conditions between species. Only the data from Reish et al. (1976) were used for
C. capitata, only data from Martin et al. (1981) and Nelson et al. (1988) were used for M. edulis, only
data from Sullivan et al. (1983) were used for E. offinis, only data from Cripe (1994) were used for P.
duorarum, and only data from Park et al. (1994) were used for Rivulus marmoratus to avoid using test
results from studies in which the life stage tested is known to be less sensitive or hi which the life stage
tested is unreported and the higher LC50s may be due primarily to the use of less sensitive life stages.
The sensitivities of the four most sensitive genera differed by a factor of 2.7, which includes two
mysids, the striped bass and the American lobster (Table 3b).
The saltwater Final Acute Value for total cadmium calculated from the Genus Mean Acute
Values in Table 3b is 80.55 ^g/L. This Final Acute Value is below the SMAV for the mysid,
Mysidopsis bigelowi (110 //g/L), but is aproximately three percent above the American lobster (78
/ig/L), approximately seven percent higher than the striped bass (75.0 Mg/L), and approximately 95
percent above the SMAV for the mysid, A. bahia (41.29 ^g/L, geometric mean of two flow-through
measured tests). The resultant saltwater Criterion Maximum Concentration (CMC) for total cadmium
is 40 jtig/L (FAV/2 or 80.55 ^g/L/2). If the total cadmium CMC is converted to dissolved cadmium
using the 0.994 factor determined experimentally by U.S. EPA, the saltwater CMC for dissolved
cadmium is 40 yC/g/L (Table 3d). The resultant 40 /zg/L CMC for dissolved cadmium is below all of the
11
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saltwater SMAVs presented in Table 3a (Figure 3).
CHRONIC TOXKTTY TO FRESHWATER ANIMALS
Acceptable chronic toxicity tests have been conducted on cadmium hi freshwater with 21
species, including seven invertebrates and 14 fishes hi 16 genera (Table 2a). Several related values are
in Table 6a. Among the unused values hi Table 6a, a 21-day Daphnia magna test hi which the test
concentrations were not measured, Biesinger and Christensen (1972) found a 16 percent reduction hi
reproduction at 0.17 /ig/L. Bertram and Hart (1979) and Ingersoll and Winner (1982) found chronic
toxicity to Daphnia pulex at less than 1 and 10 //g/L, respectively. A 32-day flow-through measured
juvenile bluegill study conducted by Cope et al. (1994) determined a growth NOEC value of >32.3
/zg/L (Table 6a), which supports the 49.8 A*g/L chronic value (Table 2a) reported by Eaton (1974).
The 200-hr LC10 of 0.7 ^g/L obtained with rainbow trout (Table 6a) by Chapman (1978) probably
would be close to the result of an early life-stage test because of the extent to which various life stages
were investigated. Effects on other salmonids and many invertebrates have been observed at 5 /ug/L
(adjusted for hardness when available) or less (Table 6a). These invertebrate species include
protozoans (Fernandez-Leborans and Noville-Villajos 1993; Niederlehner et al. 1985), C. dubia
(Winner 1988; Zuiderveen and Birge 1997), D. magna (Enserink et al. 1993; Winner and Whitford
1987), zooplankton (Lawrence and Holoka 1987), amphipods (Borgmann et al. 1991; Phipps et al.
1995), midges (Anderson et al. 1980), and mayflies (Spehar et al. 1978).
An acceptable C. dubia seven-day static-renewal toxicity test was conducted by Jop et al.
(1995) using reconstituted soft laboratory water. The < 24-hr old neonates were exposed to 1, 5, 10,
19 and 41 fJ.g/L measured cadmium concentrations hi addition to a laboratory water control at 25°C.
The NOEC and LOEC were 10 and 19 jWg/L cadmium, respectively, with a resultant chronic value of
13.78 ^g/L cadmium (Table 2a).
The effects of water hardness on the toxicity of cadmium to D. magna was evaluated by
Chapman et al. (Manuscript) under static-renewal conditions at a temperature of 20 + 2°C. As part of
the experimental design, the total hardness level was adjusted to either 53, 103 or 209 mg/L (as CaCO3)
hi three distinct tests. Daphnids were individually exposed to six measured cadmium concentrations
(exposures ranged from 0.15 to 22.1 Mg/L cadmium among the three tests) and a control (0.08 /ug/L
cadmium) for 21 days. Based on an analysis of variance hypothesis testing procedure, they reported
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reproductive (mean number of young per adult) chronic values of 0.1523, 0.2117 and 0.4371 /j.g/L
cadmium at hardness levels of 53, 103 arid 209 mg/L, respectively (Table 2a). These same data were
also subjected to a regression analysis procedure, whereby the 20 percent reproductive (mean number
of young per adult) inhibition concentration (IC20) was estimated for each hardness level. The
resultant IC20 values were 0.07, 0.23 and 0.33 //g/L cadmium for the 53, 103 and 209 mg/L hardness
levels, respectively. Overall, the results obtained by the two different procedures are similar.
The effect of cadmium on the reproduction strategy of D. magna was investigated by Bodar et
al. (1988b). After a 25-day exposure of the 12 ± 12-hr old neonates to 0 (control), 0.5, 1.0, 5.0, 10.0,
20.0 and 50 ^g/L cadmium at 20 ± 1 °C, the authors compared the survival, number of neonates per
female, first day of reproduction and neonate size of the cadmium exposures to the controls. The 25-
day reproductive NOEC was 5.0 ^g/L cadmium, and the reproductive LOEC was 10.0 /zg/L cadmium.
The resultant chronic value was 7.07 jWg/L cadmium (Table 2a).
Borgman et al. (1989) also investigated the effect of cadmium on D. magna reproduction. The
21-day static-renewal test was conducted at 20°C using measured exposure concentrations of 0.22
(control), 1.86, 4.10, 7.78 and 22.9 /^g/L cadmium. Reproduction was significantly reduced at the
lowest measured exposure concentration of 1.86 //g/L cadmium. Thus, the reproductive NOEC and
LOEC were < 1.86 and 1.86 yUg/L cadmium, respectively, with a chronic value of < 1.86 Aig/L
cadmium (Table 2a).
Brown et al. (1994) exposed 270-day old rainbow trout to cadmium under flow-through
conditions for 65 weeks using borehole water with a total hardness of 250 mg/L (as CaCO3). Mean
cadmium concentrations during the exposure of adult fish were 0.47 (control), 1.77, 3.39 and 5.48
/zg/L. After 65 weeks of exposure, the three most mature males and females were selected from each
treatment, anesthetized and striped of their gametes when possible, with the milt and ova combined in a
bucket. The fertilized eggs from each treatment group were then divided into four approximately
equal-sized subsamples and exposed for seven weeks hi 30-liter aquaria under flow-through conditions
to nominal concentrations of 0 (control), 2.0, 5.0 and 8.0 ywg/L cadmium. Second generation fry
development was significantly affected when the parents were exposed to 1.77 yWg/L cadmium, but not
when exposed to 0.47 //g/L cadmium (control). However, second generation embryo survival for all
groups was less than 60 percent, which may have influenced the fry development effect levels. A more
representative endpoint was the ability of the first generation adults to reach sexual maturity, with
NOEC and LOEC values of 3.39 and 5.48 ftg/L cadmium, respectively. The resultant chronic value
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was 4.310 fj.g/L cadmium (Table 2a).
Brown et al. (1994) also exposed two-year old brown trout to cadmium under flow-through
conditions for 95 weeks using the same borehole water. Mean cadmium concentrations during the
exposure of adult fish were 0.27 (control), 5.13, 9.34 and 29.1 //g/L. After 60 weeks of exposure, the
three most mature males and females were selected from each treatment, anesthetized and striped of
their gametes, with the milt and ova combined in a bucket. The fertilized eggs from each treatment
group were then divided into four approximately equal-sized subsamples and exposed for 50 days hi 30-
liter aquaria under flow-through conditions to cadmium concentrations similar to those in which the
parents were exposed. After the 90 week exposure, the survival NOEC and LOEC were 9.34 and 29.1
jug/L cadmium, respectively, with a resultant chronic value of 16.49 f^-gfL cadmium (Table 2a).
A 32-day fathead minnow early life stage toxicity test was conducted by Spehar and Fiandt
(1986) under flow-through conditions using sand filtered Lake Superior dilution water (Table 2a). They
reported a chronic value of 10.0 /ng/L cadmium, which when coupled with their 96-hour LC50 of 13.2
/tg/L cadmium, gives an acute-chronic ratio of 1.320.
Ingersoll and Kemble (unpublished) investigated the chronic toxicity of cadmium to the
amphipod Hyalella azteca. The organisms were exposed under flow-through measured conditions
(control, low, middle and high exposures) at a mean temperature of 23 °C and a total hardness of 280
mg/L (as CaCO3). A 3-m nylon mesh substrate was provided during the test. The seven- to eight-day
old amphipods were exposed to water only mean total cadmium concentrations of 0.10 (control), 0.12,
0.32, 0.51, 1.9 and 3.2 jag/L for 42 days. The most sensitive endpoint was survival, with an NOEC
and LOEC of 0.51 and 1.9 yug/L cadmium, respectively, after both 28 and 42 days of exposure. The
resultant chronic value was 0.9844 /^g/L total cadmium (Table 2a), which was similar to the estimated
42-day survival IC25 value of 1.9 /ug/L.
Ingersoll and Kemble (unpublished) also exposed the midge Chironomus tertians to cadmium
under the same conditions listed above for the amphipod, except that a thin 5 mm layer of sand was
provided as a substrate. The < 24-hr old larvae were exposed to water only mean measured total
cadmium concentrations of 0.15 (control), 0.50, 1.5, 3.1, 5.8 and 17.4 v-g/L for 20 days. The mean
weight, biomass, percent emergence and percent hatch endpoints all had 20-day NOEC and LOEC
values of 5.8 and 17.4 /^g/L cadmium, respectively (Table 2a). The resultant chronic value was 10.05
jUg/L total cadmium. The data were also subjected to regression analysis with resultant IC25 values of
10.3, 10.7, 8.3 and 4.0 ng/L for weight, biomass, percent emergence and percent hatch, respectively.
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All four IC25 values were similar to the 10.05 /j-g/L chronic value determined for each endpoint.
Hardness Correction
Chronic values are available over a wide range of hardness for three species (Tables 2a and
2d). To account for the apparent relationship of cadmium chronic toxicity to hardness, an analysis of
covariance (same as the analysis performed on the acute data) was performed to calculate the pooled
slope for hardness using the natural logarithm of the chronic value as the dependent variable, species as
the treatment or grouping variable, and the natural logarithm of hardness as the covariate or
independent variable. This analysis of covariance model was fit to the data in Table 2a for the three
species for which definitive chronic values are available over a range of hardness such that the highest
hardness is at least three times the lowest, and the highest is also at least 100 mg/L higher than the
lowest (other species in Table 2a did not meet these criteria). The slopes for the three species ranged
from 0.5212 to 1.579, and the pooled slope for these three species was 0.9685 with P=0.90 (Table 2c).
As with the acute slope determination, the all D. magna data set was too divergent, and only the
Chapman et al. (Manuscript) D. magna data were used with the two other species (brown trout and
fathead minnow) to estimate the overall slope. If this reduced data set is used (all species but using
only data from Chapman et al. (Manuscript) for D. magna), the pooled slope for these species was
0.7409 with P=0.35 (see Table 2c). A plot of the chronic effect level versus total hardness is provided
in Figure 4.
Criteria Development
The slope of 0.7409 was used to adjust each chronic value to a hardness of 50 mg/L.
Generally, replicate adjusted chronic values for a species agreed well, as did values for species within a
genus. The two values for Atlantic salmon are very different, but one agrees well with the value for
the other tested species in the same genus. Twenty-one Species Mean Chronic Values (SMCV) were
then calculated from the underlined values in Table 2a. When both early life stage (ELS) and life cycle
(LC) data were available for a species, the SMCV was calculated using only the LC data per the
Guideline recommendations. From these 21 SMCVs, sixteen Genus Mean Chronic Values were
calculated and ranked (Table 3c).
A freshwater Final Chronic Value was calculated from the sixteen Genus Mean Chronic Values
using the procedure used to calculate a Final Acute Value. This approach was appropriate since a
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number of chronic tests have been conducted with a large variety of species and these species met the
eight different taxonomic family Guideline requirement. Thus, the freshwater Final Chronic Value for
total cadmium at a hardness of 50 mg/L is (in Mg/L) = e^-7409^*31^68^-4.?^)^ Qr equal tQ Q 16 ^g/L
For dissolved cadmium, the Final Chronic value at a hardness of 50 mg/L is (in Mg/L) = 0.938
[e(0-7409[ln(hardness)H-719)], or equal to 0.15 Mg/L. The equation for the chronic freshwater conversion
factor is CF = 1.101672 - [(to hardness) (0.041838)] where the (hi hardness) is the natural logarithm of
the hardness (Stephen 1995). At a hardness of 50 mg/L, all Genus Mean Chronic Values are above the
dissolved Final Chronic Value (Figure 5).
Another option for calculating the Final Chronic Value is to use the Final Acute-Chronic Ratio
in conjunction with the Final Acute Value. However, the acute-chronic ratios ranged from 0.9021 for
the Chinook salmon to 433.8 for the flagfish (greater than a factor of ten), with other values scattered
throughout this range (Tables 2e and 3c). These ratios do not seem to follow any of the patterns (Table
3c) recommended in the Guidelines, and so it does not seem reasonable to use a freshwater Final Acute-
Chronic Ratio to calculate a Final Chronic Value.
CHRONIC TOXICrTY TO SALTWATER ANIMALS
Three chronic toxicity tests have been conducted with the saltwater invertebrate, Americamysis
bahia, formerly classified as Mysidopsis bahia (Table 2b). Nimmo et al. (1977a) conducted a 23-day
life-cycle test at 20 to 28°C and salinity of 15 to 23 g/kg. Survival was 10 percent at 10.6 Mg/L, 84
percent at the next lower test concentration of 6.4 Mg/L, and 95 percent hi the controls. No
unacceptable effects were observed at 6.4 Mg/L or any lower concentration. The chronic toxicity
limits, therefore, are 6.4 and 10.6 Mg/L, with a chronic value of 8.237 Mg/L. The 96-hr LC50 was
15.5 Mg/L. resulting in an acute-chronic ratio of 1.882.
Another life-cycle test was conducted on cadmium with A. bahia under different environmental
conditions, including a constant temperature of 21 °C and salinity of 30 g/kg (Gentile et al. 1982;
Lussier et al. 1985). All organisms died in 28 days at 23 Mg/L. At 10 Mg/L a series of morphological
abberations occurred at the onset of sexual maturity. External genitalia in males were abberant,
females failed to develop brood pouches, and both sexes developed a carapace malformation that
prohibited molting after the release of the initial brood. Although initial reproduction at this
concentration was successful, successive broods could not be born because molting resulted in death.
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No malformations or effects on initial or successive reproductive processes were noted in the controls
or at 5.1 /j.g/L. Thus, the chronic limits for this study are 5. kand 10 ^g/L for a chronic value of 7.141
//g/L (Table 2b). The LC50 at 21 °C and salinity of 30 g/kg was 110 jug/L which results in an acute-
chronic ratio of 15.40 from this study.
These two studies showed excellent agreement between the chronic values but considerable
divergence between the acute values and acute-chronic ratios. Several studies have demonstrated an
increase in acute toxicity of cadmium with decreasing salinity and increasing temperature (Table 6b).
The observed differences in acute toxicity to the mysids might be explained on this basis. Nimmo et al.
(1977a) conducted their acute test at 20 to 28°C and salinity of 15 to 23 g/kg, whereas the other test
was performed at 21 °C and salinity of 30 g/kg.
A third A. bahia chronic study was conducted by Carr et al. (1985) at a salinity of 30 g/kg, but
the temperature varied from 14 to 26°C over the 33 day study (Table 2b). At test termination, >50
percent of the organisms had died in cadmium exposures ^8 Atg/L. After 18 days of exposure, growth
in the 4 /^g/L, the lowest concentration treatment group was significantly reduced when compared to the
controls. The resultant chronic limits for this study are <4 and 4 Aig/L cadmium. Acute data were not
presented by the authors. The lower chronic value observed for this study as compared to the two
studies described above may have been due to unexpected temperature fluctuations over the study
period (due to mechanical problems).
Gentile et al. (1982) also conducted a life-cycle test with another mysid, Mysidopsis bigelowi,
and the results were very similar to those for A. bahia. Thus, the chronic value was 7.141 ^tg/L and
the acute-chronic ratio was 15.40.
Because they covered such a wide range, it would be inappropriate to use any of the available
freshwater acute-chronic ratios in the calculation of the saltwater Final Chronic Value. The two
saltwater species for which acute-chronic ratios are available (Table 3b) have Species Mean Acute
Values in the same range as the saltwater Final Acute Value, and so it seems reasonable to use the
geometric mean of these two ratios. When the saltwater Final Acute Value of 80.55 ^g/L is divided by
the mean acute-chronic ratio of 9.106, a saltwater Final Chronic Value of 8.9 ^g/L is obtained. The
dissolved cadmium FCV is computed using the CF (0.994 x 8.846 (j.gfL), and is equal to 8.8
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TOXICITY TO AQUATIC PLANTS
Thirty-three acceptable tests are available with freshwater plant species exposed to cadmium
which lasted from 4 to 28 days (Table 4a). Growth reduction was the major toxic effect observed with
freshwater aquatic plants, and several values are hi the range of concentrations causing chronic effects
on animals. The influence that plant growth media might have had on the toxicity tests is unknown, but
is probably minor at least hi the case of Conway (1978) who used a medium patterned after natural
Lake Michigan water. The freshwater plant and animal data presented hi this document were compared
and the lowest toxicity values for fish and invertebrate species are lower than the lowest values for
plants. A plot of the freshwater plant values is provided hi Figure 6a. Thus, water quality criteria
which protect freshwater animals should also protect freshwater plants. A final plant value was not
calculated.
Toxicity values are available for five species of saltwater diatoms and two species of
macroalgae (Table 4b). Concentrations causing fifty percent reductions hi the growth rates of diatoms
range from 60 jwg/L for Ditylum brightwelli to 22,390 /^g/L for Phaeodactylum tricornutum, the most
resistant to cadmium. The brown macroalga (kelp) exhibited mid-range sensitivity to cadmium, with an
EC50 of 860 Aig/L. The most sensitive saltwater plant tested was the red alga, Champia parvula, with
significant reductions hi the growth of both the tetrasporophyte plant and female plant occurring at 22. 8
iUg/L. The saltwater plant and animal data were also compared, and the most sensitive plant species (C.
parvula) is more resistant than the chronically most sensitive animal species tested. A plot of the
saltwater plant values is provided hi Figure 7. Therefore, water quality criteria for cadmium that
protect saltwater animals should also protect saltwater plants. A final plant value was not calculated.
BIOACCUMULATION
Bioconcentration factors (BCFs) for cadmium hi freshwater (Table 5a) range from 3 for brook
trout muscle (Benoit et al. 1976) to 6,910 for the soft tissue of the snail Viviparus georgianus (Tessier
et al. 1994b). Usually, fish accumulate only small amounts of cadmium hi muscle as compared to most
other tissues and organs (Benoit et al. 1976; Jarvinen and Ankley 1999; Sangalang and Freeman 1979).
However, specific studies summarized by Jarvinen and Ankley (1999) showed that the skin, spleen,
gill, fin, otolith and bone also have low bioconcentration factors. Sangalang and Freeman (1979) found
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that cadmium residues in fish reach steady-state only after exposure periods greatly exceeding 28 days.
D. magna, and presumably other invertebrates of about this sife or smaller, often reach steady-state
within a few days (Poldoski 1979). Cadmium accumulated by fish from water is eliminated slowly
(Benoit et al. 1976: Kumada et al. 1980), but Kumada et al. (1980) found that cadmium accumulated
from food is eliminated much more rapidly. If all variables, except temperature, were kept the same,
Tessier et al. (1994a) found that increased exposure temperatures generally increased the soft tissue
bioconcentration factor observed for the snail, V. georgianus, but not for the mussel, Elliptio
complanata. Poldoski (1979) reported that humic acid decreased the uptake of cadmium by D. magna,
but Winner (1984) did not find any effect. Ramamoorthy and Blumhagen (1984) reported that fulvic
and humic acids increased uptake of cadmium by rainbow trout.
The only BCF reported for a saltwater fish is a value of 48 from a 21-day exposure of the
mummichog (Table 6b). However, among ten species of invertebrates, the BCFs range from 22 to
3,160 for whole body and from 5 to 2,040 for muscle (Table 5b). The highest BCF was reported for
the polychaete, Ophryotrocha diadema (Klockner 1979). Although a BCF of 3,160 was attained after
sixty-four days exposure using the renewal technique, tissue residues had not reached steady-state.
BCFs for four species of saltwater bivalve molluscs range from 113 for the blue mussel
(George and Coombs 1977) to 2,150 for the eastern oyster (Zaroogian and Cheer 1976). In addition,
the range of reported BCFs is rather large for some individual species. BCFs for the oyster include 149
and 677 (Table 6b), as well as 1,220, 1,830 and 2,150 (Table 5b). Similarly, two studies with the bay
scallop resulted in BCFs of 168 (Eisler et al. 1972) and 2,040 (Pesch and Stewart 1980) and three
studies with the blue mussel reported BCFs of 113, 306, and 710 (Tables 5b and 6b). George and
Coombs (1977) studied the importance of metal speciation on cadmium accumulation in the soft tissues
ofMytilus edulis. Cadmium complexed as Cd-EDTA, Cd-alginate, Cd-humate, and Cd-pectate (Table
6b) was bioconcentrated at twice the rate of inorganic cadmium (Table 5b). Because bivalve molluscs
usually do not reach steady-state, comparisons between species may be difficult and the length of
exposure may be the major determinant in the size of the BCF.
BCFs for five species of saltwater crustaceans range from 22 to 307 for whole body and from 5
to 25 for muscle,(Tables 5b and 6b). Nimmo et al. (1977b) reported whole-body BCFs of 203 and 307
for two species of grass shrimp, Palaemonetes pugio and P. vulgaris. Vernberg et al. (1977) reported a
factor of 140 for P. pugio at 25°C (Table 6b), whereas Pesch and Stewart (1980) reported a BCF of 22
for the same species exposed at 10°C, indicating that temperature might be an important variable. The
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commercially important crustaceans, the pink shrimp and lobster, were not effective bioaccumulators of
cadmium with factors of 57 for whole body and 25 for muscle, respectively (Tables 5b and 6b).
Mallard ducks are a native wildlife species whose chronic sensitivity to cadmium has been
studied. These birds can be expected to ingest many of the freshwater and saltwater plants and animals
listed in Tables 4a and 4b. White and Finley (1978a,b) and White et al. (1978) found significant
damage at a cadmium concentration of 200 mg/kg in food for 90 days. Di Giulio and Scanlon (1984)
found significant effects on energy metabolism at 450 mg/kg, but not at 150 mg/kg. These are
concentrations which would cause damage to mallard ducks. More recent information may be
available, but these data would not have been identified during the literature search conducted for this
update.
The bioaccumulation data provided in this document is for information purposes only.
Calculation of a Final Residue Value for cadmium will not be presented at this time.
OTHER DATA
Data presented in Table 6 are not acceptable for inclusion in Tables 1-5, but provide useful
information on the effects of cadmium to aquatic organisms. Several studies were reported hi Table 6
and not in Table 1 either because the organisms were fed during acute studies (Lewis and Horning
1991; Ingersoll and Winner 1982; Mount and Norberg 1984; Pascoe et al. 1986; Schubauer-Berigan et
al. 1993; Williams and Dusenbery 1990; Wiliams et al. 1986; Winner 1984) or the tests used unusual
or uncharacterized dilution water (Hall et al. 1986; Hickey and Vickers 1992; Khangarot and Ray
1989a).
Although a number of the values hi Tables 6a and 6b have already been discussed, the
following section presents information supporting data presented in Tables 1-5, plus other useful trends
or relationships. The effects of prior cadmium exposure to the resistence of the marine copepod,
Acartia clausi, was investigated by Moraitou-Apostolopoulou et al. (1979). They observed that an A.
clausi population collected from a metal impacted area displayed a greater tolerance to lethal cadmium
concentrations when compared to a population obtained from a non-polluted site. The pollution
acclimated population also had greater longevity than the non-adapted population when exposed to
sublethal levels of cadmium.
The cumulative mortality resulting from exposure to cadmium for more than 96 hours is clearly
evident from the studies with phytoplankton (Fargasova 1993; Findlay et al. 1996), duckweed (Outridge
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1992), protozoa (Niederlehner et al. 1985), zooplankton (Lawrence and Holoka (1987), snails (Spehar
et al. 1978), zebra mussels (Kraak et al. 1992a,b), crayfish (Thorp et al. 1979), macroinvertebrates
(Giesy et al. 1979), polychaetes (Reish et al. 1976), bivalve molluscs, crabs, and starfish (Eisler and
Hennekey 1977), scallops, shrimp, and crabs (Pesch and Stewart 1980), and a mysid (Gentile et al.
1982; Nimmo et al. 1977a).
In unmeasured flow-through sockeye salmon cadmium exposures, Servizi and Martens (1978)
reported 7-day LC50 values that ranged from 8 to 4,500 //g/L for fry and alevins, respectively. The
range and life stage sensitivity pattern observed by the authors were similar to other salmonid studies
reported in Table la.
Nimmo et al. (1977a) hi studies with the mysid, Americamysis bahia, reported a 96-hr LG50 of
15.5 ^g/L (Table 1) and a 17-day LC50 of 11 (tg/L (Table 6) at 25 to 28°C and salinity of 10 to 17
g/kg hi the 96-hr study and 15 to 23 g/kg hi the 17-day study. In another series of studies with this
mysid (Gentile et al. 1982), the 96-hr LC50 was 110 ^g/L (Table 1) and the 16-day LC50 was 28 /zg/L
(Table 6b) at 20°C and salinity of 30 g/kg. These data suggest that short-term acute toxicity might be
strongly influenced by environmental variables, whereas long-term effects, even mortality, are not.
Considerable information exists concerning the effect of salinity and temperature on the acute
toxicity of cadmium. Unfortunately, the conditions and durations of exposure are so different that
adjustment of acute toxicity data for salinity is not possible. Rosenberg and Costlow (1976) studied the
synergistic effects of cadmium and salinity combined with constant and cycling temperatures on the
larval development of two estuarine crab species. They reported reduction hi survival and significant
delay hi development of the blue crab with decreasing salinity. Cadmium was three times as toxic at a
salinity of 10 g/kg than at 30 g/kg. Studies with the mud crab resulted in a similar cadmium-salinity
response. In addition, the authors report that cycling temperature may have a stimulating effect on
survival of larvae compared to constant temperature.
Theede et al. (1979) investigated the effect of temperature and salinity on the acute toxicity of
cadmium to the colonial hydroid, Laomedea loveni. At 17.5 °C cadmium concentrations inducing
irreversible retraction of half of the polyps ranged from 12.4 //g/L at a salinity of 25 g/kg to 3.0 //g/L
at 10 g/kg (Table 6). At a temperature of 17.5°C, the toxicity of cadmium increased as salinity
decreased from 25 g/kg to 10 g/kg.
A similar acute toxicity-salinity relationship was observed by Hall et al. (1995) for the copepod,
Euryfemora affinis, whereby the 96-hour toxicity increased four-fold (from 213 to 51.6 Mg/L cadmium)
when the salinity was decreased from 15 to 5 g/kg at a test temperature of 25°C. Hall et al. (1995) also
21
image:
observed an approximate three-fold toxicity increase to the sheepshead minnow when the salinity was
lowered in similar fashion at the same temperature. Likewise, the 21-day toxicity of cadmium to the
blue crab, Callinectes sapidus, increased over nine-fold when the salinity was lowered from 25 to 2.5
g/kg, and the temperature was held constant at 22-23°C (Guerin and Stickle 1995). In contrast, Snell
and Personne (1989b) observed little difference in the 24-hour toxicity of cadmium to the rotifer,
Brachiomis plicatilis, exposed under 15 and 30 g/kg salinity regimes and a temperature of 25 °C.
The effect of environmental factors on the acute toxicity of cadmium is also evident from tests
with the early life stages of saltwater vertebrates. Alderdice et al. (1979a,b,c) reported that salinity
influenced the effects of cadmium on the volume, capsule strength, and osmotic response of embryos of
the Pacific herring. Studies with embryos of the winter flounder indicated a quadratic salinity-cadmium
relationship (Voyer et al. 1977), whereas Voyer et al. (1979) reported a linear relationship between
salinity and cadmium toxicity to Atlantic silverside embryos.
Several studies have reported chronic sublethal effects of cadmium on saltwater fishes (Table
6b). Significant reduction in gill tissue respiratory rate was reported for the cunner after a 30-day
exposure to 50 ^g/L (Maclnnes et al. 1977). Dawson et al. (1977) also reported a significant decrease
in gill-tissue respiration of striped bass at 0.5 Mg/L above ambient levels after a 30-day, but not a 90-
day, exposure. A similar study with the whiter flounder (Calabrese et al. 1975) demonstrated a
significant alteration in gill tissue respiration rate measured in vitro after a 60-day exposure to 5 /
UNUSED DATA
Based on the requirements set forth hi the guidelines (Stephen et al. 1985), the following studies
are not acceptable for the following reasons and are classified as unused data.
Studies Were Conducted with Species That Are Not Resident in North America
Abbasi and Soni (1986)
Abel and Papoutsoglou (1986)
Abel and Gamer (1986)
Abel and Barlocher (1988)
Ahsanullah et al. (1981)
Ahsanullah and Williams (1991)
Amiard-Triquet et al. (1987)
Annune et al. (1994)
Arshaduddin et al. (1989)
Austen et al. (1997)
Aveiy et al. (1996)
Azeez and Baneqee (1987)
Baby and Menon (1987)
Bambang et al. (1994)
Bednarz and Warkowska-Dratnal
(1983/1984)
Birmelin et al. (1995)
Bresler and Yanko (1995)
Brooks et al. (1996)
Brunetti et al. (1991)
Calevro et al. (1998)
22
image:
Canli and Furness (1993, 1995)
Cassini et al. (1986)
Castille and Lawrence (1981)
Centeno et al. (1993)
Chan (1988)
Chandini (1988, 1988, 1989, 1991)
Chandra and Garg (1992)
Charpentier et al. (1987)
Chattopadhyay et al. (1995)
Cheung and Lam (1998)
Coppellotti (1994)
D'Agostino and Finney (1974)
Dallinger et al. (1989)
Darmono (1990)
Darmono et al. (1990)
Datta et al. (1987)
Demon et al. (1989)
Den Besten et al. (1989, 1991)
De Nicola Giudici and Guarino (1989)
De Nicola Giudici'and Migliore (1988)
Denton and Burdon-Jones (1986 1986)
Devi (1987, 1996)
Devi and Rao (1989)
Devineau and Triquet (1985)
Dorgelo et al. (1995)
Douben (1989)
Drbal et al. (1985)
Duquesne and Coll (1995)
Evtushenko et al. (1986)
Evtushenko et al. (1990)
Ferrari et al. (1993)
Fisher et al. (1996)
Fisher et al. (1996)
Forget et al. (1998)
Francesconi (1989)
Francesconi et al. (1994)
Forbes (1991)
Gaur et al. (1994)
Gerhardt (1992, 1995)
Ghosh and Chakrabarti (1990)
Glynn(1996) ^ 4,
Glynn et al. (1992, 1994) :
Gopal and Devi (1991)
Green et al. (1986)
Greenwood and Fielder (1983)
Gupta and Rajbanshi (1991)
Gupta et al. (1992)
Hader et al. (1997)
Hansten et al. (1996)
Heinis et al. (1990)
Herkovits and Coll (1993)
Hiraoka et al. (1985)
Hu et al. (1996)
Huebner and Pynnonen (1992)
Husaini et al. (1991)
Ikuta (1987)
Jenkins and Sanders (1985)
Karlsson-Norrgren and Runn (1985)
Kasuga (1980)
Keduo et al. (1987)
Khangarot and Ray. (1987)
Khristoforova et al. (1984)
Kobayashi (1971)
Krassoi and Julli (1994)
Krishnaja et al. (1987)
Kuhn and Pattard (1990)
Kuroshima (1987)
Kuroshima and Kimura (1990)
Kuroshima et al. (1993)
Lam (1996, 1996)
Lam et al. (1997)
Lee and Xu (1984)
Loumbourdis et al. (1999)
McCahon et al. (1988)
McCahon and Pascoe (1988, 1988,
1988)
McCahon et al. (1989)
McClurg (1984)
Ma et al. (1999)
Malea (1994)
Markich and Jeffree (1994, 1994)
Martinez et al. (1996)
Metayer et al. (1982)
Michibata et al. (1986)
Michibata et al. (1987)
Migliore and Giudici (1987)
Moller et al. (1994)
Mostafa and Khalil (1986)
Muino et al. (1990)
Musko et al. (1990)
Nakagawa and Ishio (1988, 1989)
Nassiri et al. (1997)
Negilski (1976)
Nir et al. (1990)
Noraho and Gaur (1995)
Notenboom et al. (1992)
Nott and Nicolaidou (1994)
Nugegoda and Rainbow (1995)
Ojaveer et al. (1980)
Pantani et al. (1997)
Papathanassiou (1995)
Pavicic et al. (1994)
Perez-Coll and Herkovits (1996)
Pynnonen (1995)
Rainbow and Kwan (1995)
Rainbow et al. (1980)
Rainbow and White (1989)
Ralph and Burchett (1998)
Ramachandran et al. (1997)
Rao and Madhyastha (1987)
Rebhun and Ben-Amotz (1984)
Reish et al. (1988)
Ringwood (1990, 1992)
Ritterhoffetal. (1996)
Romeo and Gnassia-Barelli (1995)
Safadi(1998)
Sastry and Shukla (1994)
23
image:
Sastry and Sunita (1982)
Saxenaetal. (1990,1993)
Schaferetal. (1994)
Sehgal and Saxena (1987)
Shanmukhappa and Neelakantan (1990)
Shivaraj and Patil (1988)
Simoes Goncalves (1989)
Stuhlbacher and Malfby (1992)
Takamura et al. (1989) -
Temara et al. (1996a,b)
Ten Hoopen et al. (1985)
Thaker and Haritos (1989)
Thebaultetal. (1996)
Theedeetal. (1979)
Tomasik et al. (1995)
Tyurin and Khristoforova (1993)
Udoidiong and Akpan (1991)
Valencia et al. (1998)
Van Gemert (1985)
Vashchenko and Zhadan (1993)
Verriopoulos and Moraitou-
Apostolopoulou (1981, 1982)
VisviM and Rachlin (1991)
Vogiatzis and Loumbourdis (1998)
Vrankenetal. (1985)
Vuori(1994)
Vymazal (1990, 1995)
Walsh et al. (1995)
Warnau et al. (1995a,b,c, 1996a,b,
1997)
Westernhagen and Dethlefsen (1975)
Westernhagen et al. (1975, 1978)
Wildgust and Jones (1998)
White and Rainbow (1986)
Wicklund and Runn (1988)
Wicklund et al. (1988)
Wu et al. (1997)
Wundram et al. (1996)
Zanders an&Rojas (1992, 1996)
Zou and Bu (1994)
Brown and Ahsanullah (1971) conducted tests with a brine shrimp species, that are too atypical to be
used in deriving national criteria.
Cadmium Was a Component of a Drilling Mud, Effluent, Mixture, Sediment or Sludge
Allen (1994, 1995)
Amiard-Triquet et al. (1988)
Andres et al. (1999)
Arnac and Lassus (1985)
Austen and McEvoy (1997)
Bartsch et al. (1999)
Beiras et al. (1998)
Bendell-Young (1994)
Bendell-Young etal. (1986)
Besser and Rabeni (1987)
Biesinger et al. (1986)
Bigelow and Lasenby (1991)
Bodar et al. (1990)
Buckley et al. (1985)
Burden and Bird (1994)
Busch et al. (1998)
Campbell and Evans (1991)
Camusso et al. (1995)
Carlisle and Clements (1999)
Casini and Depledge (1997)
Cuvin-Aralar (1994)
Cuvin-Aralar and Aralar (1993)
Dallinger et al. (1997)
de March (1988)
Elliott et al. (1986)
Farag et al. (1994, 1998)
Gully and Mason (1993)
Hall et al. (1984, 1987, 1988)
Hardy and Raber (1985)
Hare etal. 1991, (1994)
Haritonidis et al. (1994)
Hartwell (1997)
Haynes et al. (1989)
Hendriks (1995)
Hickey and Clements (1998)
Hickey and Martin (1995)
Hickey and Roper (1992)
Hogstrand et al. (1991)
Hollis et al. (1996)
Hooten and Carr (1998)
Hylland et al. (1996)
Inza et al. (1998)
Jak et al. (1996)
Janssens de Bisthoven et al. (1992)
Jop (1991)
Keenan and Alikhan (1991)
Kelly and Whitton (1989)
Kettle and deNoyelles (1986)
Khan and Weis (1993)
Khan et al. (1989)
Kiffhey and Clements (1996)
24
image:
Klerks and Bartholomew (1991)
Kock et al. (1995)
Koivisto et al. (1997)
Kolok et al. (1998)
Kraak et al. (1993, 1994)
Krantzberg (1989a,b)
Krantzberg and Stokes (1988, 1989)
Kumar (1991)
Lee and Luoma (1998)
Lithner et al. (1995)
Lucker et al. (1997)
Macdonald and Sprague (1988)
Maloney (1996)
Manz et al. (1994)
Marr et al. (1995a, b)
Mathew and Menon (1992)
Mersch et al. (1996)
Nalewajko (1995)
Nelson (1994) ,; i
Odin et al. (1996, 1997)
Palawski et al. (1985)
Pedersen and Petersen (1996)
Pellegrini et al. (1993)
Playleetal. (1993)
Polar and Kucukcezzar (1986)
Poulton et al. (1995)
Prevot and Soyer-Gobillard (1986)
Qichen et al. (1988)
Rachlin and Grosso (1993)
Reynoldson et al. (1996)
Richelle et al. (1995)
Roch and McCarter (1984)
Roesijadi and Fellingham (1987)
Sanchiz et al. (1999)
Schaeffer et al. (1991)
Smokorowski et al. (1997)
Stephenson and Macki (1989)
Stern and Stern (1980)
Talbot (1985, 1987)
Tessier et al. (1993)
Vuori (1993)
Vymazal (1984)
Wall et al. (1996)
Walsh and Hunter (1992)
Wang et al. (1996)
Warren et al. (1998)
Weimin et al. (1994)
Wong etal. (1982)
Woodling (1993)
Woodward et al. (1995)
These Reviews Only Contain Data That Have Been Published Elsewhere
Barnthouse et al. (1987)
Bay et al. (1993)
Cairns et al. (1985)
Chapman et al. (1968)
Dierickx and Bredael-Rozen (1996)
Dyer et al. (1997)
Eisler (1981)
Bisler et al. (1979)
Enserink et al. (1991)
Florence et al. (1992)
Guilhermino et al. (1997)
Hare (1992)
Hornstrom (1990)
Jonnalagadda and Rao (1993)
Khangarot and Ray (1987)
Kooijman and Bedaux (1996)
Kraak et al. (1994a,b)
LeBlanc (1984)
Mark and Solbe (1998)
Meyer (1999)
Nendza et al. (1997)
Oikari et al. (1992)
Papoutsoglou and Abel (1993)
Pesonen and Andersson (1997)
Phillips and Russo (1978)
Ramesha et al. (1996)
Rice (1984)
Skowronski et al. (1998)
Spry and Wiener (1991)
Thomann et al. (1997)
Thompson et al. (1972)
Toussaint et al. (1995)
Trevors et al. (1986)
Van Leeuwen et al. (1987)
Vymazal (1990)
Wright and Welbourn (1994)
Wong (1987)
25
image:
Organisms Were Exposed to Cadmium in Food or by Injection or Gavage
Bodaretal. (1988)
Brouwer et al. (1992)
Chou et al. (1986)
Daviesetal. (1997)
Decho and Luoma (1994)
Gottofrey and Tjalve (1991)
Handy (1993)
Kluttgen and Ratte (1994)
Kuroshima (1992)
Lasenby and Van Duyn (1992)
Lawrence and Holoka (1991)
Lomagin and Ul'yanova (1993)
Malley and Chang (1991)
Melgar et al. (1997)
Mount et al. (1994)
Munger and Hare (1997)
Postma et al. (1994)
Postma and Davids (1995)
Reinfelder and Fisher (1994, 1994)
Reddy et al. (1997)
Rhodes et al. (1985)
Van den Hurk et al. (1998)
Wallace and Lopez (1997)
Wang and Fisher (1996)
Wen-Xiong and Fisher (1996)
Wong (1989)
No Interpretable Concentration, Time, Response Data or Examined Only a Single Concentration
Berglind (1985)
Bitton et al. (1994)
Block and Part (1992)
Block et al. (1991)
Blondin et al. (1989)
Bowen and Engel (1996)
Bressan and Brunetti (1988)
Castano et al. (1996)
Christoffers and Ernst (1983)
Clausen et al. (1993)
Fargasova (1994)
Femandez-Pinas et al. (1995)
George et al. (1983)
Iftode et al. (1985)
Dangovan et al. (1998)
Issa et al. (1995)
Jana and Sahana (1988)
Kluytmans et al. (1988)
Kraak et al. (1993b)
Kosakowska et al. (1988)
Lussier et al. (1999)
Mateo et al. (1993)
Palackova et al. (1994)
Pereira et al. (1993)
Prasad et al. (1998)
Rachlin and Grosso (1991)
Reader et al. (1989)
Reddy and Fingennan (1994)
Reid and McDonald (1991)
Ribo(1997),
Rombough (1985)
Rosas and Ramirez (1993)
Sauvant et al. (1997)
Skowronski et al. (1991)
Sunila and Lindstrom (1985)
Trehan and Maneesha (1994)
Verbost et al. (1987)
Visviki and Rachlin (1994)
Wang et al. (1995)
Woodall et al. (1988)
Wundram et al. (1996)
Xue and Sigg (1998)
No Useable Data on Cadmium Toxicity or Bioconcentration
Battaglini et al. (1993)
Borchardt (1983)
Craig etal. (1998)
Gargiulo et al. (1996)
Gomot (1998)
Harvey and Luoma (1985)
Kraal et al. (1995)
Penttinen et al. (1995)
Rouleau et al. (1998)
Sobhan and Stemberg (1999)
26
image:
Organisms Were Selected, Adapted or Acclimated for Increased Resistance to Cadmium
Anadu et al. (1989)
Bodar et al. (1990)
Currie et al. (1998)
Ramo et al. (1987)
Herkovits and Perez-Coll (1995)
Kaplan et al. (1995)
McNicol and Scherer (1993)
Madoni et al. (1994)
Nagel and Voigt (1995)
Thomas et al. (1985)
Van Steveninck et al. (1992)
Data were not used if the results were only presented graphically (Laegreild et al. 1983; Laube
1980; Remacle et al. 1982), if the organisms were not exposed to cadmium in water (Foster 1982; >
Hatakeyama and Yasuno 1981a; O'Neill 1981), or if there was no pertinent adverse effect (Carr and
Neff 1982; DeFilippis et al. 1981; Dickson et al. 1982; Fisher and Fabris 1982; Fisher and Jones 1981;
Tucker and Matte 1980; Watling 1981; Weis et al. 1981).
Either the Materials, Methods or Results Were Insufficiently Described
Abbasi and Soni (1989)
Ball (1967)
Belabed et al. (1994)
Bendell-Young (1999)
Bitton et al. (1995)
Bjerregaard and Depledge (1994)
Bolanos et al. (1992)
Burnison et al. (1975)
Calevro et al. (1998)
Canton and Slooff (1979)
Carpene and Boni (1992)
D'Aniello et al. (1990)
Davies et al. (1994)
Department of the Environment (1973)
Errecalde et al. (1998)
Fennikoh et al. (1978)
Femandez-Leborans and Antonio-
Garcia (1988)
Galic and Sipos (1987)
Glubokov (1990)
Gorman and Skogerboe (1987)
Guanzon et al. (1994)
Guerin et al. (1994)
Hofslagare et al. (1985)
Janssen and Persoone (1993)
Jaworska et al. (1997)
Kay et al. (1986)
Kessler (1985)
Khangarot et al. (1987)
Koyama et al. (1992)
Landner and Jernelov (1969)
Lee and Oshima (1998)
Liao and Hsieh (1990)
Maas (1978)
Mansour (1993)
Ministry of Technology (1967)
Moza et al. (1995)
Munger et al. (1999)
Naylor et al. (1992)
Nwadukwe and Erondu (1996)
Pascoe and Shazili (1986)
Pauli and Berger (1997)
Penttinen et al. (1998)
Peterson (1991)
Peterson et al. (1984)
Rayms-Keller et al. (1998)
Rombough (1985)
Sandau et al. (1996)
Sekkat et al. (1992)
Shcherban (1977)
Sheela et al. (1995)
Sovenyi and Szakolczai (1993)
Stom and Zubareva (1994)
Stubblefield et al. (1999)
Tarzwell and Henderson (1960)
Verma et al. (1980)
Vykusova and Svobodova (1987)
Ward (1986)
Witeska et al. (1995)
Yamamoto and Inque (1985)
Zhang et al. (1992)
27
image:
High control mortalities occurred in testing reported by Asato and Reish (1988), Hong and
Reish (1987), Sauter et al. (1976) and Wright (1988). The 96-hr values reported by Buikema et al.
(1974a,b) were subject to error because of possible reproductive interactions (Buikema et al. 1977).
Bringmann and Kuhn (1982) and Dave et al. (1981) cultured daphnids in one water and tested them in a
different water. The acceptability of the dilution water or medium used hi some studies (e.g., Brkovic-
Popovic and Popovic 1977a,b; Cearley and Coleman 1973, 1974; Nasu et al. 1983) was open to
question because of its origin or content.
Inappropriate Medium or Medium Contained Too Much of a Complexing Agent for Algal Studies
Baillieul and Blust (1999)
Brand etal. (1986)
Cheaetal. (1997)
Couillard(1989)
Hockett and Mount (1996)
Huebertetal. (1993)
Huebert and Shay (1991, 1992, 1993)
Jenkins and Mason (1988)
Jenkins and Sanders (1986)
Jenner and Janssen-Mommen (1993)
Kessler (1986)
Lue-Kim et al. (1980)
Macfie et al. (1994)
Meteyer et al. (1988)
Muller and Payer (1979)
Nasu et al. (1988)
Rebhun and Ben-Amotz (1986, 1988)
Stary and Kratzer (1982)
Stary et al. (1983)
Sloofetal. (1995)
Sunda and Huntsman (1996)
Thongra-af and Matsuda (1993)
Thorpe and Costlow (1989)
Tortell and Price (1996)
Vasseur and Pandard (1988)
Wright et al. (1985)
Questionable Treatment of Test Organisms or Inappropriate Test Conditions or Methodology
Babich and Stotsky (1982)
Brown et al. (1984)
Bryan (1971)
Chan etal. (1981)
Dorfinan (1977)
Eisler and Gardner (1973)
Greig (1979)
Hung (1982)
Hutcheson (1975)
Moraitou-Apostolopoulou et al. (1979)
Parker (1984)
Pecon and Powell (1981)
Rehwoldt et al. (1972, 1973)
Ridlington et al. (1981)
Servizi and Martens (1978)
Sunda et al. (1978)
Wikfors and Ukeles (1982)
Bioconcentration Studies Conducted in Distilled Water, Not Conducted Long Enough,
Not Flow-through or Water Concentrations Not Adequately Measured
Allen (1995)
Amiardetal. (1993)
Amiard-Triquet et al. (1986)
Balogh and Salanki (1984)
Baudrimont et al. (1997)
Beattie and Pascoe (1978)
Bentley (1991)
Berglind (1986)
Bernds (1998)
Bervoets et al. (1995, 1996)
Bjerregaard (1982, 1985, 1991)
Block and Glynn (1992)
28
image:
Brown et al. (1986)
Burrell and Weihs (1983)
Carmichael and Fowler (1981)
Carr and Neff (1982)
Chan et al. (1992)
Chander et al. (1991)
Chawla et al. (1991)
Chitguppa et al. (1997)
Chou and Uthe (1991)
Collard and Matagne (1994)
Craig et al. (1999)
Davies et al. (1981)
De Conto Cinier et al. (1997)
De Conto Cinier et al. (1998)
De Nicola et al. (1993)
Denton and Burdon-Jones (1981)
Elliott et al. (1985)
Engel (1999)
Everaarts (1990)
Fair and Sick (1983)
Frazier and George (1983)
Freeman (1978, 1980)
Giles (1988)
Gottofrey et al. (1988)
Graney et al. (1984)
Gupta and Devi (1993)
Haines and Brumbaugh (1994)
Hansen et al. (1995)
Hardy and O'Keeffe (1985)
Hashim et al. (1997)
Hatakeyama (1987)
Herwig et al. (1989)
Hollis et al. (1997)
Irato and Piccinni (1996)
John et al. (1987)
Katti and Sathyanesan (1985)
Kerfoot and Jacobs (1976) «.
Khoshmanesh et al. (1996, t
Klaverkamp and Duncan (1987)
Koelmans et al. (1996)
Kohler and Riisgard (1982)
Kwan and Smith (1991)
Langston and Zhou (1987)
Les and Walker (1984)
McLeese and Ray (1984)
Maeda et al. (1990)
Malley et al. (1989)
Maranhao et al. (1999)
Merschetal. (1993)
Mizutani et al. (1991)
Muramoto (1980)
Mwangi and Alikhan (1993)
Nolan and Duke (1983)
Norey et al. (1990)
Oakley et al. (1983)
Olesen and Weeks (1994)
Papathanassiou (1986)
Pawlik and Skowronski (1994).
Pawlik et al. (1993)
Pelgrom et al. (1994)
Pelgrom et al. (1997)
Playle and Dixon (1993)
Presing et al. (1993)
Postma et al. (1996)
Poulsen et al. (1982)
Rai et al. 1995
Rainbow (1985)
Ramirez et al. (1989)
Ray et al. (1981)
Reichert et al. (1979)
Reinfelder et al. (1997)
Riisgard et al. (1987)
Ringwood (1989, 1992, 1993)
Roseman et al. (1994)
Rubinstein et al. (1983)
Santojanni et al. (1998)
Sedlaceketal. (1989)
Sidoumou et al. (1997)
Simoes Goncalves et al. (1988)
Sinha et al. (1994)
Skowronski and Przytocka-Jusiak
(1986)
Srivastava and Appenroth (1995)
Stary et al. (1982)
Sunil et al. (1995)
Suzuki et al. (1987)
Swinehart (1990)
Taylor et al. (1988)
Tessier et al. (1996)
Thomas et al. (1983)
Van Leeuwen et al. (1985)
Van Ginneken et al. (1999)
Vymazal (1995)
Wang and Fisher (1998)
Wading (1983a)
White and Rainbow (1982)
Williams et al. (1998)
Windom et al. (1982)
Winner and Gauss (1986)
Winter (1996)
Woodworth and Pascoe (1983)
Xiaorong et al. (1997)
Yager and Harry (1964)
Zauke et al. (1995)
Zia and McDonald (1994)
The bioconcentration tests of Eisler (1974), Jennings and Rainbow (1979b), O'Hara (1973b),
Phelps (1979), and Sick and Baptist (1979), which used radioactive isotopes of cadmium, were not used
29
image:
because of the possibility of isotope discrimination. Reports on the concentrations of cadmium in wild
aquatic organisms, such as Anderson et al. (1978), Bouquegneau and Martoja (1982), Boyden (1977),
Bryan et al. (1983), Frazier (1979), Gordon et al. (1980), Greig and Wenzloff (1978), Hazen and Kneip
(1980), Kneip and Hazen (1979), McLeese et al. (1981), Noel-Lambot et al. (1980),.Pennington et al.
(1982), Ray et al. (1981), Smith et al. (1981), and Uthe et al. (1982) were not used for the calculation
of bioaccumulation factors due to an insufficient number of measurements of the concentration of
cadmium in the water.
SUMMARY
Freshwater Species Mean Acute Values (SMAV) for cadmium are available for species in 55
genera and hardness adjusted values range from 1.613 /zg/L for brown trout to 96,880 ^g/L for a
midge. Freshwater invertebrate SMAVs range from 13.41 yug/L for D. magna to 96,880 yUg/L for a
midge and SMAVs for 24 fish species from 1.613 ,ug/L for the brown trout to 10,663 /^g/L for the
tilapia. The antagonistic effect of hardness on acute, toxicity has been demonstrated with 12 species.
Acceptable chronic tests have been conducted on cadmium with 14 freshwater fish species and seven
invertebrate species with hardness adjusted Species Mean Chronic Values (SMCV) ranging from
0.2747 jwg/L for Hyalella azteca. to 27.17 /ig/L for Ceriodaphnia dubia. Acute-chronic ratios are
available for six species and range from 0.9021 for the chinook salmon to 433.8 for the flagfish.
Freshwater aquatic plants are affected by cadmium at concentrations ranging from 2 to 20,000
/zg/L. These values are in the same range as the acute toxicity values for fish and invertebrate species,
and are considerably above the chronic values. Bioconcentration factors (BCFs) for cadmium in
freshwater range from 7 to 6,910 for invertebrates and from 3 to 2,213 for fishes.
Saltwater cadmium SMAVs are available for species in 54 genera and SMAVs for 50 species of
invertebrates range from 41.29 ^g/L for a mysid to 135,000 Mg/L for an oligochaete worm. SMAVs
for 11 fish species range from 75.0 /ug/L for striped bass to 50,000 /ug/L for sheepshead minnow. The
acute toxicity of cadmium generally increases as salinity decreases. The effect of temperature seems to
be species-specific. Chronic tests have been conducted with two mysid species, Americamysis bahia
and Mysidopsis bigelowi, with SMCVs of 6.173 Mg/L and 7.141 /ug/L, respectively. Acute-chronic
ratios are available for each species, 5.384 for A. bahia and 15.40 for M. bigelowi. The acute values
appear to reflect effects of varying salinity and temperature levels, whereas the few available chronic
values apparently do not.
Studies with macroalgae and microalgae revealed effects at 22.8 to 22,390 /-ig/L, respectively.
30
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These values are in the same range as acute toxicity values for fish and invertebrate species, and are
above the chronic values. BCFs determined with a varietyjof saltwater invertebrates range from 5 to
3,160. BCFs for bivalve molluscs were generally above 1,000 in long exposures, with no indication
that steady-state had been reached.
A comparison of the criteria developed in this document with the previous National
recommended water quality criteria (which is based on the 1995 update for freshwater and the 1984
update for saltwater) indicates that the updated 2001 freshwater CMC of 1.0 //g/L dissolved cadmium
has remained approximately the same (the value was lowered each time to protect the commercially
important rainbow trout), but the freshwater chronic CCC has been lowered to 0.15 //g/L dissolved
cadmium hi this document from 1.3 yWg/L in the 1995 document. This 2001 update contains a database
of 55 freshwater genera for acute toxicity (43 genera were in the 1995 update), and 15 genera for
freshwater chronic toxicity (12 genera were provided in the 1995 document). As a result of the
additional data, the acute and chronic hardness derived slopes are different hi this update relative to
previous versions. This update did not use an adjusted "n" value to calculate the Final Chronic Value
(the 1995 update modified the total "n" for the chronic value to be the same as the acute "n" value).
Included hi this updated document are toxicity results for certain threatened and endangered species that
were not available earlier. Saltwater cadmium criteria remained relatively the same between the 1999
National recommended water quality criteria and 2001 documents. The new saltwater CMC of 40 /t/g/L
dissolved cadmium presented in this document is only slightly lower than the 42 Mg/L cadmium found
hi the previous national recommended water quality criteria. The chronic CCC dropped slightly to 8.8
/zg/L cadmium hi this document from the 9.3 //g/L value previously recommended. There are 54
genera hi the acute saltwater database of this document (the 1984 document had 33 genera), and the
same two saltwater chronic genera are presented hi both documents (a third A. bahia chronic value was
added to this document).
NATIONAL CRITERIA
The available toxicity data, when evaluated using the procedures described hi the "Guidelines
for Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important species is unusually sensitive,
freshwater aquatic life should be protected at a total hardness of 50 mg/L as CaCO3 if the four-day
average concentration (in ^g/L) of dissolved cadmium does not exceed the numerical value given by
31
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0.938 [6 ' P11^ ess*< • •>] more man once every ^QQ years on me average, and if the 24-hour
average dissolved concentration (in ^g/L) does not exceed the numerical value given by 0.973
[6 ] more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaCO3 the four-day average dissolved concentrations of
cadmium are 0.15, 0.25 and 0.40 ^g/L, respectively, and the 24-hour average dissolved concentrations
are 1.0, 2.0, and 3.9 A*g/L.
The procedures described in the "Guidelines for Deriving Numerical National Water Quality
Criteria for the Protection of Aquatic Organisms and Their Uses" indicate that, except possibly where a
locally important species is unusually sensitive, saltwater aquatic life should be protected if the four-day
average dissolved concentration of cadmium does not exceed 8.8 //g/L more than once every three
years on the average and if the 24-hour average dissolved concentration does not exceed 40 yMg/L more
than once every three years on the average. However, the limited data suggest that the acute toxicity of
cadmium is salinity-dependent; therefore the 24-hour average concentration might be underprotective at
low salinities and overprotective at high salinities.
U.S. EPA believes that the use of dissolved cadmium will provide a more scientifically correct
basis upon which to establish water-column criteria for metals. The criteria were developed on this
basis. The use of dissolved criteria reduces the amount of conservatism that was present in earlier
cadmium criteria. It is recognized that a considerable proportion of dissolved cadmium in organic-rich
waters may be less toxic than freely dissolved cadmium. On the other hand, some paniculate forms of
cadmium might contribute to cadmium loading of organisms, possibly through ingestion.
A return interval of three years continues to be the Agency's general recommendation. The
resilience of ecosystems and their ability to recover differ greatly, however, and site-specific criteria
may be established if adequate justification is provided.
The use of criteria in designing waste treatment facilities requires the selection of an
appropriate wasteload allocation model. Dynamic models are preferred for the application of these
criteria. Limited data or other factors may make their use unpractical, in which case one should rely
on a steady-state model. The Agency recommends the interim use of 1Q5 or 1Q10 for Criterion
Maximum Concentration (CMC) design flow and 7Q5 or 7Q10 for the Criterion Continuous
Concentration (CCC) design flow in steady-state models for unstressed and stressed systems
respectively. These matters are discussed hi more detail hi the Technical Support Document for Water
Quality-Based Toxics Control (U.S. EPA 1991).
32
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