Cadmium Risks to Freshwater Life: 

Derivation and Validation of Low-Effect 

Criteria Values using Laboratory and 

Field Studies

Scientific Investigations Report 2006–5245

Version 1.2, September 2010

Prepared in cooperation with the Idaho Department of Environmental Quality

U.S. Department of the Interior

U.S. Geological Survey
Cover: The montage illustrates an analytical approach and photographs of species that were important 

inthis report as follows:

Graphic: A species-sensitivity distribution (SSD) of freshwater species to chronic cadmium exposures 

is illustrated. The SSD concept and analyses were fundamental to this report. The open circles and 

horizontal lines represent the means and ranges of available chronic effects thresholds to cadmium 

for different species, respectively. Selected species are labeled.

Photographs:

Left: The amphipod Hyalella azteca; photograph by Scott Bauer, U.S. Department of Agriculture, 

Agricultural Research Service.

Top right: A westslope cutthroat trout, Oncorhynchus clarki lewisi; photograph by Ernest R. Keeley, Idaho 

State University.

Middle right: A mottled sculpin, Cottus bairdi, from captive broodstock that was used for toxicity testing; 

photograph by Doug Hardesty, U.S. Geological Survey, Columbia Ecological Research Center, 

Columbia, Missouri.

Bottom right: A mottled sculpin in a natural setting; photograph by Ernest R. Keeley, Idaho State University.
Cadmium Risks to Freshwater Life: 

Derivation and Validation of Low-Effect 

Criteria Values using Laboratory and 

Field Studies

By Christopher A. Mebane

Prepared in cooperation with the Idaho Department of Environmental Quality

Scientific Investigations Report 2006–5245

Version 1.2, September 2010

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

DIRK KEMPTHORNE, Secretary

U.S. Geological Survey

Mark D. Myers Director

U.S. Geological Survey, Reston, Virginia: 2006

Revised September 2010

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U.S. Government.

Although this report is in the public domain, permission must be secured from the individual copyright owners to 

reproduce any copyrighted materials contained within this report.

Suggested citation:

Mebane, C.A., 2006 (2010 rev.), Cadmium risks to freshwater life: Derivation and validation of low-effect criteria 

values using laboratory and field studies (version 1.2): U.S. Geological Survey Scientific Investigations Report 

2006-5245, 130 p.
iii

Acknowledgments

Many people generously shared their time and insights to provide data, explanations, or 

helpful criticisms. Reviewers included Harry Ohlendorf, CH2M Hill, Sacramento, Calif.; Steven 

Canton, Chadwick Ecological Consultants, Littleton, Colo.; Charles Stephan, U.S. Environmental 

Protection Agency (EPA) Environmental Research Laboratory, Duluth, Minn.; Don Essig and 

Joanna Sandow, Idaho Department of Environmental Quality, Boise, Idaho; Christopher Ingersoll, 

U.S. Geological Survey, Columbia Ecological Research Center, Columbia, Mo.; and Daren 

Carlisle, U.S. Geological Survey National Water Quality Assessment, Reston, Va.

Many investigators generously took time to answer questions, provide additional details of 

their work, share data, or suggest other information sources. These included John Besser, 

Nile Kemble, and Stacy James, U.S. Geological Survey Columbia Ecological Research Center, 

Columbia, Mo.; Stephen Brinkman, Colorado Division of Wildlife, Fort Collins, Colo.; Uwe 

Borgmann, National Water Research Institute, Burlington, Ontario, Canada; Victor Castillo 

III, Southwest Texas State University, San Marcos, Tex.; Michaël Coeurdassier, University of 

Franche-Comte, Besançon, France; Valery Forbes, Roskilde University, Roskilde, Denmark; 

Clyde Goulden, National Academy of Science Philadelphia; Duane Gossiaux, National Oceanic 

and Atmospheric Administration, Ann Arbor, Mich.; Lydia Hollis, Golder Associates, Calgary; 

David Mount, U.S. Environmental Protection Agency, Environmental Research Laboratory, 

Duluth, Minn.; David Nimick, U.S. Geological Survey, Helena, Mont., Vince Palace and Richard 

McNicol, Freshwater Research Institute, Winnipeg; David Riddell, University of New Brunswick, 

Fredericton, New Brunswick, Canada; Andrew Wigginton, University of Kentucky, Lexington, Ky.; 

and Chris Wood, McMaster University, Waterloo, Ontario, Canada. Russell Erickson, 

U.S. Environmental Protection Agency, Environmental Research Laboratory, Duluth, Minn., 

gave statistical advice; Anne Kuhn and Glenn Thursby, U.S. Environmental Protection Agency, 

Environmental Research Laboratory, Narragansett, R.I., and Webb Van Winkle, Oak Ridge 

National Laboratory (retired), Boise, Idaho provided advice on population modeling. The Idaho 

Department of Environmental Quality’s Surface Water Program provided principal funding for 

this project. Don A. Essig was the project manager for the Idaho Department of Environmental 

Quality.
iv

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v

Contents

Acknowledgments........................................................................................................................................iii

Abstract...........................................................................................................................................................1

Introduction.....................................................................................................................................................2

Cadmium in Ambient Waters ..............................................................................................................2

Methods...........................................................................................................................................................3

Methods for Criteria Derivation..........................................................................................................3

Overview of Criteria Derivation Steps......................................................................................3

Extrapolating Small Toxicity Test Datasets to Aquatic Communities .................................4

Minimum Species-Diversity Required for Criteria Derivation..............................................8

Important Elements and Assumptions of Criteria Derivation ...............................................8

Data Sources and Inclusion Requirements ..........................................................................10

Methods for Evaluation of Risks to Biological Integrity and Vulnerable Species under 

Criteria Conditions.................................................................................................................11

Results ...........................................................................................................................................................11

Criteria Derivation Results ................................................................................................................12

Data Compilation and Evaluation ............................................................................................12

Static, Renewal, and Flow-Through Exposures...........................................................12

Tests with No Analytical Confirmation of Exposure Concentrations .......................13

Life-Cycle Versus Shorter-Term Chronic Data.............................................................14

Statistical Interpretation of Chronic Tests....................................................................15

Diversity of Data.........................................................................................................................17

Acute Toxicity of Cadmium to Freshwater Animals ............................................................17

Chronic Toxicity of Cadmium to Freshwater Animals ..........................................................17

Other Data on the Effects of Cadmium to Freshwater Animals. ........................................18

Hardness–Toxicity Relations....................................................................................................19

Acute to Chronic Ratios............................................................................................................22

Hardness-Based Cadmium Criteria Equations .....................................................................24

Aquatic Toxicology of Cadmium—Synthesis .................................................................................30

Mechanisms and Factors Affecting Aquatic Toxicology of Cadmium..............................30

Toxicity of Cadmium in Chemical Mixtures...................................................................31

Relative Sensitivity of Chronic Test Endpoints.............................................................31

Bioaccumulation and Effects of Dietary Exposures to Cadmium.............................32

Fish .......................................................................................................................32

Aquatic Invertebrates .............................................................................................37

Bioconcentration and Bioaccumulation Factors ...............................................38

Behavioral Toxicity ...........................................................................................................39

Sensitivity of Different Life Stages of Aquatic Organisms to Cadmium...................41

Acclimation........................................................................................................................41
vi

Contents—Continued

Figures

Figure 1. Graph showing comparison of the assumed distributions of species mean 

acute and chronic cadmium values (triangles) for all North American species 

and the distributions of measured values ................................................ 6

Figure 2. Graphs showing hardness-toxicity relations for acute and chronic data ............ 20

Figure 3. Graphs showing variability in dissolved and chronic cadmium criterion 

concentrations in selected streams in Idaho and Montana, 1997–2005 ............ 45

Figure 4. Graphs showing risks to Hyalella azteca populations exposed to the cadmium 

chronic criterion ........................................................................... 54

Figure 5. Graphs showing species sensitivity distributions of cadmium values showing 

species mean values, 5th percentile final acute and chronic values, and their 

fitted logistic distribution curves ......................................................... 57

Figure 6. Graphs showing comparison of measured cadmium chronic criterion 

exceedence factors, and corresponding instream biological condition metrics ... 60

Figure 7. Graphs showing comparison of cadmium effects concentrations estimated from 

ecosystem studies or field assessments to proportions of adversely affected 

species based on a species-sensitivity distribution (SSD) of laboratory-derived 

species mean chronic values ............................................................ 63

Results—Continued

Evaluation of Risks to Biological Integrity and Vulnerable Species...........................................42

Criteria in Variable Environments............................................................................................42

Risks of the 5th Percentile Species-Sensitivity Distribution (SSD)-Based Chronic 

Criterion to a More Sensitive Species......................................................................48

Role of Hyalella azteca in Aquatic Food Chains ..........................................................48

Hyalella azteca in Waters with Elevated Cadmium Concentrations........................49

Simulating Effects of Cadmium to a Natural, Coldwater Hyalella azteca

Population.............................................................................................................50

Population Modeling Methods ..............................................................................51

Population Modeling Results.................................................................................53

Field Validation ...........................................................................................................................56

Cadmium Whole-Lake Ecosystem Experiment ............................................................58

Co-Occurrence of Cadmium Criterion Exceedences and Apparent Effects in 

Streams.................................................................................................................59

Cadmium and Threatened and Endangered Species ..........................................................66

Indirect Effects to Listed Salmonids Via Effects to Sensitive Prey Items ...............68

Potential Site-Specific Adjustments to Cadmium Criteria Values .....................................68

Summary........................................................................................................................................................69

References Cited..........................................................................................................................................70

Appendix – Final Review...........................................................................................................................125

Glossary.......................................................................................................................................................128
vii

Tables

Table 1. Comparison of the diversity of selected aquatic assemblages in North America

to the diversity of aquatic species with cadmium toxicity data available ............ 5

Table 2. Comparison of the minimum taxonomic diversity needed to derive aquatic-life

criteria and the diversity represented in the dataset of chronic effects of

cadmium on aquatic animals............................................................... 8

Table 3. Acute toxicity values versus hardness for individual and pooled regressions ...... 21

Table 4. Chronic toxicity values versus hardness for individual and pooled regressions ... 21

Table 5. Other data on chronic toxicity values versus hardness not included in 

individual and pooled regressions......................................................... 22

Table 6. Cadmium acute-chronic ratios (ACRs), ranked in order of closeness of the

species mean acute values (SMAV) to the final acute value (FAV) .................. 23

Table 7. Ranked genus mean acute and chronic values ....................................... 25

Table 8. Acute and chronic criteria calculations ................................................ 28

Table 9. Comparison of acute and chronic cadmium criteria values derived in this 

report to previously reported cadmium criteria.......................................... 30

Table 10. Examples of cadmium tissue residues co-occurring with the absence or

presence of adverse effects ............................................................... 33

Table 11. Effects of cadmium to the most chronically sensitive species, Hyalella azteca,

at chronic criterion concentrations ...................................................... 50

Table 12. Vital rates used for modeling effects of cadmium to a coldwater Hyalella azteca

population .................................................................................... 52

Table 13. Comparison of chronic criterion values with results of ecosystem studies 

and field surveys ........................................................................... 64

Table 14. Comparison of estimated no-effect concentrations for threatened and

endangered species occurring in Idaho to chronic criteria concentrations......... 67

Table 15. Acute toxicity of cadmium to freshwater animals ....................................... 89

Table 16. Chronic toxicity of cadmium to freshwater animals ....................................102

Table 17. Other data on effects of cadmium to freshwater organisms ...........................109
viii

Conversion Factors

[Emphasizing conversions that were used to compile the tables in this report using comparable

units of concentration. Good sources for additional characteristics of water include Hem (1992)

and Hounslow (1995). Atomic weights are from Weast (1990)]

Multiply By To obtain

microgram per kilogram (μg/kg) 0.001 milligram per kilogram

millequivalents calcium 20.04 milligrams per liter calcium

milliequivalents magnesium 12.15 milligrams per liter magnesium

milligram per kilogram (mg/kg) 1,000 microgram per kilogram

millimoles calcium (mM) 40.078 milligrams per liter calcium

millimoles magnesium (mM) 24.305 milligrams per liter magnesium

micromoles cadmium (μM) 112.411 microgram per liter

nanomoles cadmium (nM) 0.112411 micrograms per liter

milligrams per liter (mg/L) 1,000 micrograms per liter

nanograms per liter (ng/L) 0.001 micrograms per liter

water hardness as milliequivalents (meq) 50.05 hardness as calcium carbonate 

(CaCO3) in milligrams per liter

water hardness as millimoles (mM) 100.09 Hardness as calcium carbonate 

(CaCO3) in milligrams per liter

gram (g) 0.03527 ounce, avoirdupois

square meter (m2) 0.0002471 acre

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows:

°F=(1.8×°C)+32

Water hardness as milligrams per liter (mg/L) calcium carbonate may be derived from 

concentrations of calcium (Ca) and magnesium (Mg) as follows:

Hardness (as CaCO3) in mg/L) = 2.497 x (Ca in mg/L) + 4.118 x (Mg in mg/L).

Concentrations of chemical constituents in water are given either in milligrams per liter (mg/L) 

or micrograms per liter (μg/L).
Abstract

In 2001, the U.S. Environmental Protection Agency 

(EPA) released updated aquatic life criteria for cadmium. 

Since then, additional data on the effects of cadmium to 

aquatic life have become available from studies supported by 

the EPA, Idaho Department of Environmental Quality (IDEQ), 

and the U.S. Geological Survey, among other sources. Updated 

data on the effects of cadmium to aquatic life were compiled 

and reviewed and low-effect concentrations were estimated. 

Low-effect values were calculated using EPA’s guidelines 

for deriving numerical national water-quality criteria for the 

protection of aquatic organisms and their uses. Data on the 

short-term (acute) effects of cadmium on North American 

freshwater species that were suitable for criteria derivation 

were located for 69 species representing 57 genera and 33 

families. For longer-term (chronic) effects of cadmium on 

North American freshwater species, suitable data were located 

for 28 species representing 21 genera and 17 families. Both 

the acute and chronic toxicity of cadmium were dependent on 

the hardness of the test water. Hardness-toxicity regressions 

were developed for both acute and chronic datasets so that 

effects data from different tests could be adjusted to a common 

water hardness. Hardness-adjusted effects values were pooled 

to obtain species and genus mean acute and chronic values, 

which then were ranked by their sensitivity to cadmium. 

The four most sensitive genera to acute exposures were, 

in order of increasing cadmium resistance, Oncorhynchus

(Pacific trout and salmon), Salvelinus (“char” trout), Salmo

(Atlantic trout and salmon), and Cottus (sculpin). The four 

most sensitive genera to chronic exposures were Hyalella

(amphipod), Cottus, Gammarus (amphipod), and Salvelinus. 

Using the updated datasets, hardness dependent criteria 

equations were calculated for acute and chronic exposures to 

cadmium. At a hardness of 50 mg/L as calcium carbonate, the 

criterion maximum concentration (CMC, or “acute” criterion) 

was calculated as 0.75 μg/L cadmium using the hardnessdependent equation CMC = e(0.8403  ́ ln(hardness)-3.572) where the 

“ln hardness” is the natural logarithm of the water hardness. 

Likewise, the criterion continuous concentration (CCC, or 

“chronic” criterion) was calculated as 0.37 μg/L cadmium 

using the hardness-dependent equation CCC = (e(0.6247  ́

ln(hardness)-3.384))  ́ (1.101672 – ((ln hardness)  ́ 0.041838))). 

Using data that were independent of those used to 

derive the criteria, the criteria concentrations were evaluated 

to estimate whether adverse effects were expected to the 

biological integrity of natural waters or to selected species 

listed as threatened or endangered. One species was identified 

that would not be fully protected by the derived CCC, the 

amphipod Hyalella azteca. Exposure to CCC conditions likely 

would lead to population decreases in Hyalella azteca, the 

food web consequences of which probably would be slight 

if macroinvertebrate communities were otherwise diverse. 

Some data also suggested adverse behavioral changes are 

possible in fish following long-term exposures to low levels 

of cadmium, particularly in char (genus Salvelinus). Although 

ambiguous, these data indicate a need to periodically review 

the literature on behavioral changes in fish following metals 

exposure as more information becomes available. Most data 

reviewed indicated that criteria conditions were unlikely 

to contribute to overt adverse effects to either biological 

integrity or listed species. If elevated cadmium concentrations 

that approach the chronic criterion values occur in ambient 

waters, careful biological monitoring of invertebrate and fish 

assemblages would be prudent to validate the prediction that 

the assemblages would not be adversely affected by cadmium 

at criterion concentrations.

Cadmium Risks to Freshwater Life: Derivation and 

Validation of Low-Effect Criteria Values using Laboratory 

and Field Studies

By Christopher A. Mebane
Introduction

In 2001, the U.S. Environmental Protection Agency 

(EPA) published an update of aquatic life criteria for 

cadmium. Since the publication of the U.S. Environmental 

Protection Agency’s (2001) update, additional data on the 

effects of cadmium on aquatic life have been generated, 

including data from studies that focused on Idaho species or 

conditions. These include studies conducted or sponsored by 

EPA, Idaho Department of Environmental Quality (IDEQ), 

U.S. Geological Survey (USGS), and an Idaho discharger, the 

Thompson Creek Mining Company (Mebane, 2003; Chadwick 

Ecological Consultants, Inc., 2004a; Besser and others, 2007). 

In cooperation with the IDEQ, the USGS compiled and 

reviewed updated data on the effects of cadmium to aquatic 

life and estimated low-effect values. The main purposes of 

this report are (1) to compile updated datasets on the effects 

of cadmium to aquatic organisms, and to synthesize the 

compilation using EPA’s procedures for deriving numerical 

national criteria for the protection of aquatic life (Stephan 

and others, 1985); and (2) to evaluate whether the calculated 

criteria would likely protect aquatic communities under 

realistic field conditions. The second purpose emphasizes 

species and conditions that are relevant to Idaho 

Section 304(a) of the Clean Water Act (CWA, 33 U.S. 

Code §1314(a)) requires that EPA publish, and from time 

to time revise, water-quality criteria that accurately reflect 

the latest scientific knowledge on the fate and effects of 

pollutants in water bodies. Section 304(a) of the Clean Water 

Act listed numerous specific areas of scientific knowledge to 

be addressed by water-quality criteria for protecting aquatic 

life and human uses of waters. Topics specified by the statute 

pertaining to aquatic life criteria (as opposed to human 

health criteria) include (1) identifiable effects to plankton, 

fish, shellfish, wildlife, and plant life; (2) concentration 

and dispersal of pollutants through biological, physical, 

and chemical processes; and (3) the effects of pollutants on 

biological community diversity, productivity, and stability 

for varying types of receiving waters. Federal water-quality 

standards regulations further provide that States adopt numeric 

water-quality criteria based on either EPA’s criteria, EPA’s 

criteria modified to reflect site-specific conditions, or other 

scientifically defensible methods (40 CFR 131.11). 

As a “scientifically defensible method of analysis,” this 

analysis largely relies on EPA’s 1985 guidelines for developing 

numerical National aquatic-life criteria. EPA’s guidelines (“the 

guidelines”) were developed to provide an objective, internally 

consistent, appropriate, and feasible way of deriving National 

criteria (Stephan and others, 1985). Although the EPA 

distinguishes between regulatory “water-quality standards” 

and non-regulatory “aquatic-life criteria” (U.S. Environmental 

Protection Agency, 1994), the words “criterion” and “criteria” 

often imply regulatory limits in other contexts. As used here, 

an acute or a chronic criterion is a concentration and exposure 

duration corresponding to a specified risk probability and does 

not imply a regulatory limit. Synonyms to “criteria” in this 

context might include low-effect concentrations, benchmarks, 

guidelines, targets, thresholds, or risk values. 

Cadmium in Ambient Waters

Cadmium, like other heavy metals, generally is rare 

in unpolluted natural waters. The term “heavy metals” 

refers to metallic elements with an atomic number greater 

than about 20, the atomic number of calcium. Cadmium 

often co-occurs with copper and zinc. Cadmium, copper, 

and zinc concentrations in natural waters apparently are 

roughly proportional to their relative abundance in rocks. 

In granite, typical cadmium concentrations are about 

150 to 400 times less than copper or zinc concentrations, 

respectively (about 0.13, 20, and 50 mg/kg, respectively for 

cadmium, copper, and zinc). This rank order between the 

three metals’ concentrations, and the approximate two orders 

of magnitude difference between cadmium and copper or 

zinc concentrations also was reported for basalt, shale, and 

limestone (Drever, 1995). In natural waters, typical cadmium 

concentrations are about 10–50 times less than copper or 

zinc concentrations (Hem, 1992; Stephan and others, 1994a; 

Drever, 1995). 

Concentrations of trace metals are often reported as total 

or dissolved concentrations. “Total” or “total recoverable” 

concentrations usually refer to values measured from raw, 

unfiltered wholewater samples. “Dissolved” concentrations 

usually refer to values measured from water samples that 

have been physically filtered to separate particulate and 

aqueous fractions. For filtered trace-metal samples, the 

standard procedure used by the EPA and USGS is to filter 

wholewater samples through a 0.45-micrometer (μm) filter 

(U.S. Environmental Protection Agency, 1994; Wilde and 

others, 2004). In this report, when filtered (dissolved) values 

are available, they are emphasized because of the belief that 

dissolved metal more closely approximates the bioavailable 

fraction of metal in the water column than does total metal 

(Prothro, 1993). As a practical matter, with cadmium this 

distinction is less important than for some other trace metals 

such as copper or lead. Because cadmium is highly soluble in 

water, differences between unfiltered (total concentrations) 

and 0.45-μm filtered samples are usually small, with dissolved 

cadmium concentrations averaging about 90 to 95 percent 

those of unfiltered “total” samples (Stephan, 1995; Clark, 

2002). 

Cadmium concentration data collected from natural 

waters may be biased high in some older publications. As 

of 1971, the median natural background concentration of 

cadmium in waters of the United States was estimated at 

about 1 μg/L (Hem, 1992). In contrast, Stephan and others 

(1994b) estimated that typical background dissolved cadmium 

concentrations in freshwaters of the United States ranged 

2 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
from 0.002 to 0.08 μg/L. The difference in these estimates 

does not indicate that over two decades ambient cadmium 

concentrations decreased by one to two orders of magnitude, 

but rather reduced field and laboratory sample contamination 

through the use of “clean” techniques (Stephan and others, 

1994a). Even in locations in North America that because 

of unusually elevated cadmium concentrations were the 

subject of several investigations, concentrations in streams 

and lakes still seldom reach more than about 15 μg/L. These 

locations include the vicinities of Rouyn-Noranda, Quebec 

and Sudbury, Ontario, Canada, and Coeur d’Alene, Idaho 

(Clark, 2002; Couture and Kumar, 2003; this report). In Idaho, 

data collected using “clean” field and laboratory techniques 

indicate that background concentrations of dissolved cadmium 

range from <0.02 to 0.1 μg/L (see section “Criteria in Variable 

Environments”), consistent with Stephan and others (1994a) 

estimate. 

Cadmium concentrations can become elevated in waters 

that are influenced by sources such as mining, minerals 

processing, and combustion of fossil fuel. Cadmium is present 

in zinc ore minerals such as sphalerite and is recovered from 

some copper ores during smelting and refining. Cadmium 

has a tendency to enter the atmosphere through vaporization 

at high temperatures; therefore, cadmium may be released 

to the environment in metallurgical processes and fossil fuel 

combustion. Cadmium is used for electroplating, for pigments 

in paint, printing ink, and plastics, as a stabilizer for PVC 

plastics, and in electrical batteries. Many of these uses tend to 

make the element available to water that comes into contact 

with wastes (Hem, 1992). Cadmium also may be elevated in 

wastewater discharges due to corrosion of galvanized pipes in 

public water-supply systems (R. Finch, City of Boise, Idaho, 

oral commun., 2005). 

Although rare in surface waters, cadmium is highly toxic 

to some aquatic life. In comparative acute toxicity testing of 

all 63 atomically stable heavy metals in the periodic table, 

cadmium clearly was the most toxic metal (Borgmann and 

others, 2005). Recent EPA recommendations for regulatory 

criteria to protect aquatic life from cadmium toxicity are not 

much higher than the upper range estimates of background 

cadmium concentrations. U.S. Environmental Protection 

Agency (2001) recommended that the highest 4-day average 

dissolved cadmium concentration not exceed 0.08 to 0.25 μg/L 

at water hardnesses of 20 and 100 mg/L as calcium carbonate, 

respectively. 

Methods

This analysis involved two distinct parts. First, datasets 

of acute (short-term) and chronic (long-term) responses of 

aquatic organisms to cadmium were compiled and synthesized 

following EPA guidelines (Stephan and others, 1985). Second, 

the resulting criteria, field survey results, and ecosystem 

experiments were compared to evaluate the protectiveness 

of criteria in realistic settings. Criteria also were compared 

with estimated thresholds of minimal effects to especially 

vulnerable species listed as threatened or endangered under 

the Endangered Species Act.

Methods for Criteria Derivation

This analysis followed the general approach described 

in EPA’s National guidelines for deriving numerical waterquality criteria (Stephan and others, 1985). Major steps in 

the approach and deviations from the guidelines are briefly 

described here, with further details described in the results.

Overview of Criteria Derivation Steps

1. Data compilation and review—The guidelines call for 

collecting all available, relevant data on the toxicity to, 

and bioaccumulation by, aquatic animals and plants, 

Food and Drug Administration Action levels, and 

feeding studies or field studies of wildlife that regularly 

consume aquatic organisms. Literature was searched for 

relevant data through electronic databases, reviewing 

contents of selected key journals, and contacts with 

other investigators to locate unpublished data. Four 

primary electronic databases were searched. The Web of 

Science®—Science Citation Index, the Aquatic Sciences 

and Fisheries Abstracts, and Water Resources Abstracts 

(Cambridge Scientific Abstracts, 2005a, 2005b; Thomson 

Scientific, 2005) were searched using the keywords 

“cadmium” and “toxicity.” EPA’s Ecotox database was 

searched using as search criteria cadmium and any 

survival, growth, reproduction, or behavioral endpoints 

(U.S. Environmental Protection Agency, 2002a). Most 

searches were from January 1, 2000, through April 2005, 

based on the premise that the bibliography of EPA’s 

2001 review was sufficiently comprehensive for data 

published in up to and including 1999. Additionally, 

because not all authors used the term “toxicity” in their 

keywords or titles, the following journals were searched 

for the same period using “cadmium” as a search term 

in titles, keywords, or abstracts: Aquatic Toxicology, 

Canadian Journal of Fisheries and Aquatic Sciences, and 

Environmental Toxicology and Chemistry. With more 

than 2,000 search results, at least the titles were read to 

evaluate if the papers might contain useful information. 

Additionally, a relevant report was obtained and reviewed 

for additional data (Chadwick Ecological Consultants, 

Inc., 2004b). EPA’s 2001 review concluded that plants 

were less sensitive to cadmium than aquatic animals. 

EPA’s review did not include dietary effects to wildlife 

(U.S. Environmental Protection Agency, 2001). Neither 

plant toxicity or dietary effects of cadmium to wildlife 

were reviewed in this report. 

Methods 3
Data were not used if test methods were questionable or 

insufficiently described. Although Stephan and others 

(1985) provided guidelines for screening questionable 

data, other judgments of data acceptability were 

postponed until the overall dataset was compiled and 

analyzed. Guidelines for screening data include whether 

the tests included a control treatment, mortality in the 

control treatments was too high, organisms were stressed 

or previously exposed to substantial concentrations of test 

materials or other pollutants, use of deionized water or 

other inappropriate dilution water, whether the test species 

does not have reproducing populations in the wild in 

North America, or whether the organisms were exposed to 

chemicals in a mixture (Stephan and others, 1985).

2. Develop hardness-toxicity regressions —Because of 

repeated observations of a linear relation between the 

natural logarithms of hardness and acute toxicity values 

for metals (Stephan and others, 1985; Meyer, 1999), 

no effort was made to examine other possible relations, 

such as two-variable models using pH and hardness. 

Data available over a range of hardness for a species 

were useful for establishing hardness-toxicity relations 

if they included a broad range of hardness, if agreement 

among the species was reasonably good, and if hardnesstoxicity slopes between different species were reasonably 

consistent. The slopes from each species’ hardnesstoxicity relation were pooled through an analysis of 

covariance (Zar, 1984).

3. Determine genus mean acute and chronic values 

(GMAVs and GMCVs) for all test species at a standard 

hardness level and rank order the GMAVs and GMCVs 

by sensitivity—Acute values were based on short-term 

exposures (48 to 96 hours) that killed 50 percent of the 

organisms in the test (LC50). Using the hardness-toxicity 

relations, each value was adjusted to a common hardness 

value. If after hardness-normalization, values for a 

species and values that appear questionable compared 

with other values within the genus probably should not 

be used. Stephan and others (1985) gave a factor of 10 

difference as an example in which some or all values 

probably should be rejected. No fixed factor was used 

for this analysis, rather data were examined species-byspecies to judge whether values were so dissimilar to 

be questionable. For example, for a species with many 

values (for example rainbow trout with 37 values), a 

single extreme value would not change the geometric 

mean for the species much, and if no individual value was 

obviously anomalous, all data were included. In contrast, 

if only two data points were available, and differed by 

a factor of about 3, or if method or result details were 

limited or questionable, one or more values might not be 

used.

4. Calculate the final acute value (FAV) and final chronic 

value (FCV) – The FAV and FCV are acute and 

chronic effects concentrations corresponding to the 5th

percentile genus in the respective datasets. More details 

and assumptions for this step are described later in the 

“Methods” section.

5. Derive the hardness-dependent criterion maximum 

concentration (CMC or acute criterion) and the criterion 

continuous concentration (CCC or chronic criterion) 

equations using the hardness-toxicity slopes – The CMC 

is equal to the FAV divided by two. This step extrapolates 

from a concentration that would likely be extremely 

harmful to sensitive species in short-term exposures (kill 

50 percent of the population) to a concentration expected 

to kill few if any individuals. This assumption was 

recently supported by results of the acute toxicity testing 

of 20 species with 5 chemicals representing a broad range 

of toxic modes of action. In those data, multiplying the 

LC50 by a factor of 0.56 resulted in a low- or no-acute 

effect concentration (Dwyer and others, 2005b). No 

analogous adjustment was made to the FCV because it is 

derived using effects less severe than killing one-half the 

population of sensitive species.

Extrapolating Small Toxicity Test Datasets to 

Aquatic Communities 

Aquatic-life criteria are intended to apply to a diversity 

of freshwater ecosystems and the species that inhabit them. 

Conceptually, it would be desirable to develop criteria 

for substances that address the diversity of ecosystems by 

conducting a series of field experiments on a wide variety 

of unpolluted water body types (for example, lakes, rivers, 

and streams that are warm or cold, large or small). However, 

such realism is infeasible, and instead of testing diverse 

communities or assemblages of organisms, criteria are usually 

derived from the results of single species toxicity tests on 

a variety of test organisms (Stephan and other, 1985). The 

problem of extrapolating toxicity values from small datasets 

to estimate values needed to protect diverse aquatic-life 

communities is illustrated by comparing the diversity of 

aquatic life in North America and Idaho to the diversity of 

organisms that have been tested with cadmium. At least 1,000 

times more aquatic animals are native to North America than 

have been tested for their short-term (acute) sensitivity to 

cadmium (table 1). This disparity is greater for long-term 

(chronic) data because chronic tests are much more costly than 

acute tests, and because chronic tests can only be conducted 

with species that can be cultured or at least maintained in 

laboratories. 

4 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
In the United States, the approach generally has been 

to develop species-sensitivity distributions (SSDs) of acute 

toxicity. To account for untested and potentially sensitive 

taxa, a hypothetical LC50 for the 95th percentile most sensitive 

species is calculated from the SSD for a substance. This 

hypothetical LC50, referred to as the “final acute value” by 

Stephan and others, 1985), then is divided by an overall acuteto-chronic ratio to estimate a chronic criterion that would 

not result in unacceptable adverse effects. Although chronic 

criteria could be derived directly from chronic data, because 

much fewer chronic data are available than acute data, in 

practice acute-to-chronic ratios have been more commonly 

used to estimate the chronic toxicity of untested species 

and derive chronic criteria in the United States (Stephan 

and others, 1985; Mount and others, 2003). For example, 

EPA developed aquatic-life criteria for about 25 “priority 

pollutants” and several other “nonpriority” pollutants or 

characteristics including ammonia and dissolved oxygen (U.S. 

Environmental Protection Agency, 2002b). Of these, only 

the chronic criteria for four pollutants (ammonia, cadmium, 

selenium, and saltwater dissolved oxygen) were developed 

directly with chronic effects data without using an acute-tochronic extrapolation.

Extrapolating single-species toxicity values to derive 

criteria to protect ecosystems requires making some 

fundamental assumptions. These include: 

1. The shape of the statistical distribution of the sensitivities 

of untested organisms; 

2. Test data can be considered a random sample of the 

responses of all species, most of which are untested; 

3. Acute and chronic responses in laboratory reasonably 

approximate those for the same species in the wild; and

4. The overall distribution of single species responses 

approximates the distribution that would be found in 

community responses. 

A fundamental assumption for deriving aquatic-life criteria 

is that responses of a variety of individual species to a 

substance measured through single-species toxicity testing 

can be extrapolated to responses at the aquatic community 

level (Stephan and others, 1985; Erickson and Stephan, 

1988). Species sensitivities are further assumed constant 

and are defined and predicted through models based on 

statistical distributions and thus are commonly referred to as 

species-sensitivity-distributions (SSDs). SSDs are statistical 

distributions estimated from toxicity data for a dataset of 

species responses and are considered to be a sample of an 

entire population (in the statistical sense) of species responses. 

The idea of a SSD approach followed the observation that 

in multiple aquatic toxicity datasets of LC50 values, the species 

sensitivity was distributed in a rather constant way for most 

chemicals, in a distribution that resembles a lognormal one. 

Therefore, no one tested species is representative of any other 

species, but is one estimate of the general species sensitivity, 

that is, one point along the distribution (Posthuma and others, 

2002). A further implicit assumption of the SSD approach to 

setting environmental criteria or in ecological risk assessment 

is that risks cannot be completely eliminated but should be 

reduced to an acceptable low risk. Thus, SSDs are used to 

calculate the concentration at which a specified proportion 

of species will be affected, referred to as the hazardous 

concentration (HC) for the p (percentile) of species (HCp) 

(Posthuma and others, 2002). The HC used in water-quality 

criteria derivation is the HC5, where the HC5 in aquatic-life 

derivation is derived from the 5th percentile of the genus mean 

acute or chronic values (GMAVs or GMCVs). The choice of 

the 5th percentile as the HC5 value was simply based on the 

judgment that the 1st or 10th percentile would result in criteria 

that seemed too low or too high when compared with the 

datasets from which they were calculated. The 5th percentile 

was an easily recognizable number that was midway between 

the 1st and 10th percentile (Stephan and others, 1985; Stephan, 

2002). The HC5 level is the most commonly used value in 

the practice of using SSD to set environmental criteria or 

ecological risk assessment; however, the choice of this level 

probably is rooted in convention, for no clear ecological or 

toxicological reasons support the choice (Posthuma and 

others, 2002). 

The description of the use of the 5th percentile of a SSD 

to derive water-quality criteria implies that 95 percent of the 

species in an ecosystem would be protected, and 5 percent 

could be sacrificed for each chemical criterion. If important 

species had toxicity values less than the 5th percentile, then 

criteria should be set lower than the 5th percentile to protect 

Assemblage

Number of 

native species

Number of species in 

available datasets

North 

America Idaho Acute Chronic

Zooplankton 1 ~480 Unknown 5 4

Benthic macroinvertebrates

1 >9,000 2

>1,100 37 5

Fish 3

~950 4 40 23 17

Other organisms Unknown Unknown 4 2

1

Data from Thorp and Covich (2001).

2

Data from Grafe and others (2002).

3

Data from Matthews (1998).

4

Data from Zaroban and others (1999).

Table 1. Comparison of the diversity of selected aquatic 

assemblages in North America to the diversity of aquatic species 

with cadmium toxicity data available.

[Number of species in acute and chronic datasets are from this report. 

Symbols: ~, approximately; >, greater than]

Methods 5
these species (Stephan and others, 1985; Posthuma and others, 

2002). In their description of “important” species, Stephan 

and others (1985) considered only recreational or commercial 

value as reasons for considering a species important; however, 

over time descriptions of important or critical species in 

related contexts were broadened. Subsequent descriptions of 

important or critical species in the context of ecological risk 

assessment or site-specific aquatic-life criteria included four 

general categories of valued species: 

1. Keystone species of great ecological value because 

their loss would indirectly affect many other species 

or ecosystem function; 

2. Species that have been given special conservation 

status through treaties, laws, or policies as 

threatened, endangered, or vulnerable; 

3. Recreationally valued species; or 

4. Commercially important species (U.S. 

Environmental Protection Agency, 1994; Stephan 

and others, 1994b; Posthuma and others, 2002).

Erickson and Stephan (1988) developed a statistical 

procedure for estimating the 5th percentile of the SSD that was 

intended to minimize statistical bias in most datasets. Their 

procedure used extrapolation or interpolation to estimate 

the 5th percentile of a statistical population of genus mean 

values in which the available GMAVs were assumed to have 

been randomly obtained. The available GMAVs were ranked 

from low to high and the cumulative probability for each was 

calculated as P=R/(N+1), where R = rank and N = number of 

GMAVs in the dataset. The calculation used the log-triangular 

distribution and the four GMAVs whose P values were closest 

to 0.05. The calculation procedure is not dependent upon the 

taxonomic level of the dataset, for example, species, genus, 

or family (Erickson and Stephan, 1988). Although previous 

versions of EPA’s criteria derivation guidelines used species 

or family, in the 1985 version, criteria were determined 

at the genus level. On average, species within a genus are 

toxicologically more similar than species in different genera. 

The use of genus mean values is intended to prevent datasets 

from becoming biased by an overabundance of species in one 

or few genera.

Erickson and Stephan (1988) recommended a triangularshaped statistical distribution of toxicity values. Figure 1

shows the concept of a triangularly distributed “true” 

population of sensitivities for North American species, 

overlain on the available samples of acute and chronic 

sensitivities to cadmium. The triangular-shaped distribution 

may look odd compared to the more familiar illustrations of 

bell-shaped, normal-distributions with a curved hump in the 

center and the tails of the data gradually tapering towards, but 

Figure 1. Comparison of the assumed distributions of species mean acute and chronic cadmium values (triangles) for all 

North American species and the distributions of measured values (bars).

0.1

1

10

100

1,000

10,000

0.1

1

10

100

1,000

10,000

<0.25

<1<4

<16<64

<256

<1024<4096

<16,384

<0.25

<1<4

<16<64

<256

<1024<4096

<16,384

Number of species

Species Mean Acute Values Species Mean Chronic Values

Cadmium SMAV, in micrograms per liter Cadmium SMCV, in micrograms per liter

6 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
never meeting zero. However, the triangular distribution better 

fits the toxicological assumption that SSDs do have definite 

lower and upper bounds (that is, no species is sensitive to 

infinitesimal concentrations and no species is resistant to any 

concentration of a substance). Further, in aquatic-life criteria 

derivation, the primary interest is the more sensitive lower 

corner of the distribution (5th percentile), so the unrealistic 

peak at the center of the distribution is unimportant. 

An important feature of procedures that use the 5th

percentile of SSDs for deriving criteria values is that if the 

dataset contains less than 20 genus mean values, the estimate 

of the 5th percentile will be determined by extrapolation, and 

if greater than 20 values are used, the 5th percentile will be 

determined by interpolation. This is significant because if 

the dataset is small, the extrapolated estimate will be lower 

than the lowest value in the dataset. As the size of the dataset 

increases, the 5th percentile estimate usually increases, and 

uncertainties in its estimate owing to extrapolation and bias 

are reduced (Erickson and Stephan, 1988; Newman and others, 

2000).

The statistical procedure for deriving criteria values 

assumes that tested species are a random sample from the 

entire population of North American species (Erickson 

and Stephan, 1988). If this were the case, because insects 

and other invertebrates make up more than 90 percent of 

the aquatic animals in North American freshwaters, about 

90 percent of the toxicity data for cadmium would be for 

invertebrates. As shown in the section “Diversity of Data,” 

such is not the case. Species are not selected at random for 

testing, but are selected based on factors such as cost, whether 

they are considered as surrogates for untested valued species, 

sensitivity to pollutants, availability of standard methods and 

test organisms, repeatability, happenstance, and regulatory 

requirements. Erickson and Stephan (1988), recognizing that 

species selection was more haphazard than random, tested 

various simulations with extreme deviations from random 

sampling and concluded that questions about the propriety of 

applying methods based upon random sampling to datasets 

where sampling was not strictly random probably was not of 

great importance.

The assumptions and practice of using the “5th percentile 

of a SSD” approach to setting environmental quality criteria 

and assessing ecological risk have been the subjects of 

spirited debate in the ecotoxicology literature (Forbes and 

Forbes, 1993a; Smith and Cairns, 1993; Calow and others, 

1997; Power and McCarty, 1997; Aldenberg and Jaworska, 

2000; Newman and others, 2000; Forbes and Calow, 2002; 

Posthuma and others, 2002; Selk and others, 2002; Suter 

and others, 2002b; Fisher and Burton, 2003; Brix and others, 

2005; Maltby and others, 2005). The root of the debates are 

arguments that fundamental assumptions of the approach may 

not be met or are untested. Some concerns include: 

1. Whether haphazard collections of data from singlespecies laboratory toxicity tests can be considered 

representative of any natural ecosystems;

2. Whether small datasets can be significantly biased 

toward more or less sensitive species than would be 

expected in natural ecosystems; 

3. Whether any species loss from a community due to 

a toxin is acceptable. Supporting arguments are that 

the 5th percentile of a SSD is simply a pragmatic 

“statistical cutoff” to estimate a predicted low- or 

no-effect concentration to protect ecosystems, rather 

than a real sacrifice of species. Further, accepting 

some species loss, especially among “lower” 

organisms, is acceptable because ecosystems have 

enough functional redundancy to absorb species 

loss. The counterpoint to this redundant species 

hypothesis combines the “rivet popper” hypothesis 

(community integrity is reduced by each loss of a 

species) with the argument that a conservative stance 

is best when faced with uncertainty in ecosystem 

functions; 

4. Whether the 5th percentile of the SSD is the 

appropriate level of protection or just a familiar 

number; 

5. Reducing community integrity to a simple 

proportion of species could discount keystone 

or dominant species if they were in the lower 5th

percentile of sensitivity; 

6. The approach depends on comparable data which 

results in a bias toward mortality data, which are 

most abundant, and a bias against data on abnormal 

behavior or other sublethal data that may be as 

important for maintaining biological integrity; 

7. The few species for which multiple tests results 

are available sometimes show high variability in 

sensitivity. Thus, apparent differences between 

species’ ranks on a SSD may be unreliable, 

especially for species with only single or few 

datapoints. and, 

8. Statistical uncertainties. 

Because of these concerns, this report goes beyond the 

minimum steps to derive criteria values following Stephan 

and others (1985) guidelines and, as the available data allow, 

evaluated these concerns in the context of cadmium effects to 

relevant aquatic ecosystems. Evaluations included predictions 

of the consequences of not protecting specific species that 

have cadmium sensitivities in the lower 5th percentile of the 

SSD, and comparisons of the criteria derived through singlespecies testing in laboratory conditions to model ecosystem or 

field studies.

Methods 7
Minimum Species-Diversity Required for Criteria 

Derivation

The statistical desirability of selecting species at random 

to derive a statistical sample to base criteria upon is countered 

by the toxicological desirability of targeting for inclusion 

of specific groups that were considered important and were 

expected to be sensitive to many substances. Thus, Stephan 

and others (1985) recommended that the minimum data 

requirements for deriving aquatic life criteria include results 

from at least eight different families representing important 

groups of taxa (table 2). The minimum species-diversity 

guidelines were focused on acute toxicity data for developing 

a final acute value, and included procedures for extrapolating a 

chronic criterion from acute results based on acute-to-chronic 

ratios (ACRs). However, if chronic data met the minimum 

eight-family rule, as is the case for cadmium, a chronic 

criterion should be developed using directly from the chronic 

data using the same procedures that Erickson and Stephan 

(1988) developed for acute data. 

Important Elements and Assumptions of Criteria 

Derivation

Stephan and others’ (1985, p. 18) guidelines for 

deriving aquatic-life criteria provide an objective and 

rigorous analytical framework, but much of their guidance is 

necessarily qualitative rather than quantitative. They caution, 

Table 2. Comparison of the minimum taxonomic diversity needed to derive aquatic-life criteria and the diversity represented in the 

dataset of chronic effects of cadmium on aquatic animals.

[Minimum species diversity from Stephan and others (1985)]

Minimum species diversity needed to 

derive aquatic life criteria Family Species represented

A representative of:

1. The family Salmonidae in the class Osteichthyes Salmonidae Atlantic salmon, Salmo salar

Brook trout, Salvelinus fontinalis

Brown trout, Salmo trutta

Bull trout, Salvelinus confluentus

Chinook salmon, Oncorhynchus tshawytscha

Coho salmon, Oncorhynchus kisutch

Lake trout, Salvelinus namaycush

Rainbow trout, Oncorhynchus mykiss

2. A second family in the class Osteichthyes (preferably including 

important warm water species)

Catostomidae White sucker, Catostomus commersoni

Centrarchidae Lepomis macrochirus, Bluegill

Centrarchidae Smallmouth bass, Micropterus dolomieui

Cichlidae Blue tilapia, Oreochromis aurea

Cottidae Mottled sculpin, Cottus bairdi

Cyprinidae Fathead minnow, Pimephales promelas

Cyprinodontidae American flagfish, Jordanella floridae

Esocidae Northern pike, Esox lucius

Percidae Fountain darter, Etheostoma fonticola

3. A third family in the phylum Chordata (may be in the class 

Osteichthyes or an amphibian)

Bufonidae American toad, Bufo americanus

Ambystoma Northwestern salamander, Ambystoma gracile

4. A planktonic crustacean (cladoceran, copepod, etc.) Daphniidae Cladoceran, Ceriodaphnia dubia

Cladoceran, Ceriodaphnia reticulata

Cladoceran, Daphnia magna

Cladoceran, Daphnia pulex

5. A benthic crustacean (amphipod, crayfish) Gammaridae Amphipod, Gammarus fasciatus

Hyalellidae Amphipod, Hyalella azteca

6. An insect Chironomidae Midge, Chironomus tentans

7. A family in a phylum other than Arthropoda or Chordata 

(Annelida, Mollusca, etc.)

Aeolosomatidae Polychaete worm, Aelosoma headleyi

8. A family in any order of insect or any phylum not already 

represented

Physidae Mollusc, Aplexa hypnorum

8 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
“...much judgment will usually be required to derive 

a water-quality criterion for aquatic organisms 

and their uses” and that “All necessary decisions 

should be based on a thorough knowledge of aquatic 

toxicology and an understanding of these Guidelines 

and should be consistent with spirit of these 

Guidelines, i.e., to make best use of the available 

data to derive the most appropriate criteria.” 

This analysis strives to adhere to these principles. Several 

considerations from the guidelines that were important in this 

analysis, deviations from the guidelines, and assumptions are 

described below:

Focus on aquatic animals—U.S. Environmental 

Protection Agency (2001) compared the relative sensitivities 

of aquatic plants and animals to cadmium toxicity and 

concluded that criteria that adequately protect aquatic animals 

and their uses would be sufficient to protect aquatic plants as 

well. Therefore, no further investigation was done regarding 

plant toxicity. Only one of the studies compiled by U.S. 

Environmental Protection Agency (2001) reported cadmium 

effects at a concentration (2 μg/L) in the sensitivity range of 

sensitive animals, with all other plant values considerably 

higher. Also, the 2 μg/L value was from a test that did not 

meet data acceptability guidelines: cadmium concentrations 

were not measured, water hardness and the duration of the test 

were not reported.

Tissue residue values and bioaccumulation or 

bioconcentration—EPA’s criteria derivation guidelines include 

a procedure for deriving a “final residue value” to prevent 

tissue concentrations in important aquatic species from 

exceeding applicable U.S. Food and Drug Administration 

(FDA) action levels and to protect wildlife, including fishes 

and birds, that consume aquatic organisms from demonstrated 

unacceptable effects. The final residue value is defined as 

a concentration in water and is calculated by dividing a 

maximum permissible tissue concentration by appropriate 

bioconcentration or bioaccumulation factors. For wildlife, 

a maximum permissible tissue concentration could be 

established by defining a maximum acceptable dietary intake 

based on observations of survival, growth, or reproduction in 

a chronic feeding study or a long-term field study (Stephan 

and others, 1985). No regulatory action levels for cadmium 

and human health applicable to the consumption of aquatic 

organisms were located on the FDA’s websites, therefore, 

no basis exists for developing a residue-based criterion for 

cadmium based on exceeding FDA action levels. For wildlife, 

literature on effects of dietary cadmium and effects associated 

with tissue residues from field studies was evaluated. 

Direct calculation of a final chronic value is preferable 

to extrapolations using acute-to-chronic ratios—In EPA’s 

aquatic-life criteria documents, chronic criteria are most 

commonly based on the final acute value that was derived for 

acute criteria, and then using an acute-to-chronic ratio, were 

extrapolated to chronic criteria values. This extrapolation is 

often necessary because chronic data are scarcer than acute 

data. Chronic testing requires more time and resources than 

acute testing and chronic test methods have been developed 

for fewer species than have acute test methods. However, if 

sufficient chronic values are available, derivation of a chronic 

criterion directly from chronic values is considered preferable 

to extrapolating a chronic criterion from acute values. Stephan 

and others (1985) advise that if chronic values are available 

for species in at least eight specified families, the final chronic 

value should be calculated using the same statistical procedure 

as that used to develop the final acute value. Because chronic 

values were available for diverse enough species to meet 

the eight family minimum data requirements to use the FAV 

procedure with chronic data (table 2), updated ACRs and a 

final ACR (overall ACR for cadmium) are reported here, but 

were not used in the chronic criterion calculations. 

Updated ACRs may have utility should the cadmium 

criteria developed following the National procedures be 

modified to reflect site-specific conditions. Water-quality 

criteria for aquatic life developed following the National 

guidelines may be under- or over-protective at a specific 

site if the species at the site are more or less sensitive than 

those included in the dataset, or if physical and (or) chemical 

characteristics of the site alter the biological availability and 

(or) toxicity of the chemical. EPA established three procedures 

for the development of site-specific water-quality criteria. 

Each procedure uses a final ACR for a substance (U.S. 

Environmental Protection Agency, 1994):

1. The recalculation procedure is intended to take into 

account relevant differences between the sensitivities 

of the aquatic organisms in the nationally based dataset 

and the sensitivities of organisms that occur at the site. 

Organisms are added or deleted from the National dataset 

and criteria recalculated. If the eight-family diversity rule 

is not met for chronic values, chronic criteria would be 

calculated using a final ACR. 

2. The water-effect ratio (WER) procedure is intended 

to take into account relevant differences between the 

toxicities of the chemical in laboratory dilution water and 

in site water. Usually paired acute toxicity tests using 

“standard” test organisms such as the fathead minnow are 

conducted in site-water and a “typical” laboratory dilution 

water, providing a water-effect ratio of site and laboratory 

water LC50s. The site-specific criteria are derived by 

multiplying the Statewide acute criterion by the WER, 

and dividing that result by the final ACR provides a 

chronic criterion. 

3. The resident species procedure is intended to take 

into account both differences simultaneously. In this 

procedure, the acute criterion is derived for a site by 

testing at least eight species in site water and calculating 

an acute criterion using National guidelines. The chronic 

criterion may be derived using the final ACR for the 

chemical, or a site-specific ACR may be developed.

Methods 9
Data Sources and Inclusion Requirements 

Because toxicity testing can be done using various 

methods, guidelines for data comparability and inclusion 

are needed. Stephan and others (1985) offer several specific 

guidelines for data acceptability for criteria derivation. Based 

on the data reviewed herein, additional decisions were made 

regarding which results were suitable for pooling in the 

criteria derivation dataset. The term “acceptable data” as 

used in this context simply refers to whether testing methods 

were sufficiently similar to one another, that differences 

in results can be reasonably attributed to differences in 

species sensitivity, characteristics of test water, or biological 

variability rather than testing artifacts. All data were presumed 

acceptable for the purposes for which they were generated. 

General guidelines include: 

Exclusion of unwritten or otherwise questionable 

data – Data should be available in a written form, such as a 

publication, manuscript, letter, or memorandum (electronic 

mail was not widely used in 1985, but is considered 

an acceptable written form here). Confidential or other 

information not available for distribution should not be used. 

Questionable data, whether published or unpublished, should 

not be used. For example, data from tests that did not contain 

a control treatment, tests in which too many control organisms 

died or were stressed, or tests using mixtures of chemicals 

should not be used (Stephan and others, 1985). 

Resident in North America —Stephan and others (1985) 

recommend rejecting data for species with no reproducing 

wild populations in North America. For this report, this 

data-criterion was modified to rejecting families that are 

not known to have reproducing wild populations in North 

America. This modification is based on the premise that (1) 

taxa within a genus are toxicologically much more similar 

than taxa in different families, and (2) families may be more 

widely distributed than species. For example, aquatic insect 

assemblages in cool, temperate streams in North America and 

New Zealand share few if any native animal species, but share 

many families that have shown similar sensitivities to metals 

(Hickey and Golding, 2002). This residency provision also is 

problematic because introduced species are steadily becoming 

established, and arguably, invasive nuisance species need not 

be protected or considered in criteria. However, following 

the concept that each tested species is not representative of 

any other species but is one estimate of the general species 

sensitivity, no exclusions were made for species that can be 

considered invasive nuisances. 

Life stage sensitivity—Because a species can only be 

considered protected from acute toxicity if all life stages are 

protected, Stephan and others (1985) recommended that if 

the available data indicate that some life stages are at least a 

factor of two more resistant than other life stages, the data for 

the more resistant life stages should not be used to calculate 

species mean acute values. Smaller, juvenile life stages of 

fish and early invertebrate instars are commonly expected to 

be more vulnerable to metals toxicity than larger, older life 

stages of the same species. In acute testing methods, fish are 

recommended to be post-larval or older and actively feeding, 

daphnids less than 24 hours old; amphipods, mayflies, and 

stoneflies in an early instar; and midges in the second or third 

instar (American Society for Testing and Materials, 1997). 

Static versus flow-through exposure systems—For some 

highly volatile, hydrolysable, or degradable materials, it is 

probably appropriate to use only results of flow-through 

tests (tests in which the test solutions are constantly being 

replenished) in which concentrations of the test solutions were 

measured often enough using acceptable analytical methods. 

Tests with fed organisms—In most instances, acute test 

data were only used if the animals were not fed during the 

tests. Stephan and others (1985) advise that, 

“...results of acute tests during which the test 

organisms were fed should not be used, unless data 

indicate that the food did not affect the toxicity of the 

test material.” 

The reason for this guidance is that fecal matter and uneaten 

food will decrease the dissolved-oxygen concentration and 

the biological activity of some test materials. These problems 

are most severe with the static technique, but sometimes are 

important with the renewal and flow-through techniques 

(American Society for Testing and Materials, 1997).

This was primarily an issue for very small organisms 

that do not have the energy reserves to survive the customary 

96-hour duration of an acute toxicity test without feeding. 

Because of this, Stephan and others (1985) recommended 

that only the results of 48-hour toxicity tests be used to 

determine mean acute values for cladocerans and midges, 

instead of 96 hours used for other organisms. Although this 

recommendation avoids the potential problems of introducing 

organic material into test vessels, it introduces a potential 

bias toward more resistant mean acute values resulting from 

the shorter test durations. Suedel and others (1997) reported 

that LC50 values from 48-hour toxicity tests with cadmium 

and the cladocerans Ceriodaphnia dubia and Daphnia magna

and the midge Chironomus tentans were 2 to 4 times higher 

than 96-hour LC50s. Because comparisons among mean acute 

values for different taxa assume that values are unbiased, 

uncertainty is added when pooling or comparing values 

that were determined from both 48- and 96-hour tests, such 

as species-sensitivity rankings and acute to chronic ratio 

comparisons. 

The data reviewed did not consistently indicate that 

feeding would appreciably affect the toxicity of cadmium 

in acute tests. Lewis and Weber (1985) reported no effect 

of feeding on the sensitivity of Daphnia pulex to cadmium, 

but Daphnia magna were more resistant. However, feeding 

did not appear to make Daphnia magna more resistant in 

tests by Nebeker and others (1986a). Severe starvation of the 

amphipod Hyalella azteca prior to acute toxicity tests with 

cadmium did not result in lower LC50s. Starved amphipods 

10 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
exposed to low cadmium concentrations actually survived 

longer than control organisms (McNulty and others, 1999). 

Therefore, tests in which organisms were fed were evaluated 

for inclusion in this report on a case-by-case basis. 

Feeding and nutritional status of test organisms is 

important in longer exposures as well. For example, fewer 

mortalities, larger growth, and higher reproductive rates were 

observed with Hyalella azteca tested in nutrient rich effluents 

than with Hyalella azteca tested in nutrient poor reconstituted 

laboratory waters (Stanley and others, 2005). In contrast, tests 

with daphnids reared with different food rations indicated that 

cadmium exposure significantly decreased survival, growth, 

and reproduction and this decrease was more pronounced 

with increasing food concentration. Therefore, the sensitivity 

of daphnids to stress increased with increasing food ration. 

This increased sensitivity is likely the result of a change in 

life history from emphasizing survival at low food supply 

to stressing reproduction at high food supply (Smolders and 

others, 2005).

Concordance with other data for taxa —Values that 

appear to be questionable in comparison with other acute 

and chronic data available for the same species and for other 

species within the same genus probably should not be used. 

For example, if after adjustment for hardness, acute values for 

a species or genus differ by more than a factor of 10, rejection 

of some or all values probably is appropriate (Stephan and 

others, 1985). 

Methods for Evaluation of Risks to Biological 

Integrity and Vulnerable Species under Criteria 

Conditions

Following derivation of criteria values, the values were 

evaluated in the context of whether waters with cadmium 

concentrations near criteria conditions would be expected to 

maintain biological integrity. This effort included:

1. Evaluating the criteria concentrations that were 

derived from toxicity tests done in laboratories under 

constant conditions in the context of fluctuating 

cadmium concentrations in field conditions; 

2. Critically examining the assumption that the most 

sensitive 5 percent of taxa from a species-sensitivity 

distribution need not be explicitly protected in 

criteria derivation by considering the role that a 

more-sensitive “unprotected” species has in aquatic 

food webs and through population modeling; and 

3. Comparing apparent effects concentrations from 

field surveys or ecosystem studies to corresponding 

criteria concentrations. 

Derived criteria values were additionally evaluated to 

predict if they would harm or kill species listed under the 

Endangered Species Act as threatened or endangered species. 

Estimates of no-observed-adverse-effect-concentrations 

(NOECs) were made for several threatened or endangered 

species and compared to corresponding chronic criterion 

concentrations. The use of NOECs is more conservative 

than the method used to derive most “chronic values” for 

the criteria derivation. For criterion derivation, most chronic 

values were obtained by taking the geometric mean of the 

NOEC and the lowest-observed-adverse-effect concentration 

(LOEC) from a test. Because the chronic value lies between a 

no- (or presumably at least a small effect) concentration and a 

concentration that did cause adverse effects, by design chronic 

values could allow some amount of adverse effects. NOECs 

were selected as the test parameter used to make effects 

determinations of chronic criteria to listed species in EPA’s 

methods manual for conducting biological evaluations of 

water quality criteria (U.S. Environmental Protection Agency, 

2003b).

Results

Results are presented in three major interrelated parts:

1. Data compilations, data reduction, and criteria 

derivation following the steps in Stephan and others’ 

(1985) guidelines. The steps consisted of:

a. Compiling and evaluating datasets of acute and 

chronic cadmium values (tables 15 and 16, at 

back of report); 

b. Compiling a dataset of other data that were not 

directly comparable to the acute and chronic 

datasets, but were still pertinent (table 17, at 

back of report);

c. Determining hardness-toxicity relations for 

acute and chronic effects of cadmium to various 

species (tables 3-5); 

d. Calculation of acute to chronic toxicity ratios 

(table 6);

e. Ranking the sensitivities of the hardnessadjusted acute-chronic values for each genus 

(table 7);

f. Calculation of the 5th percentile of the ranked 

acute and chronic genus mean values (final 

acute value and final chronic value, respectively 

(table 8);

g. Derivation of hardness-dependent acute and 

chronic criteria equations (table 9). 

Results 11
2. An interpretive synthesis of the aquatic toxicology 

of cadmium in the context of the derived criteria, 

including a review of some effects data that could 

not be directly used in the criteria derivation. This 

synthesis includes observations on the importance 

of alternative testing methods on the acceptability 

of data. These include factors such as alternative 

exposure designs (flow-through, renewal, or static), 

whether organisms were or were not fed, life 

stage sensitivity including the influence of the life 

stage when tests began on the acclimation of later 

life stages to cadmium, duration of chronic tests, 

effects associated with bioaccumulation and dietary 

exposures to cadmium (table 10), and behavioral 

effects of cadmium. 

3. These analyses were followed by “validation” 

evaluations of the suitability of the derived cadmium 

criteria using different types of data than were used 

to develop them. These included considerations 

for applying the criteria to variable environments 

(fig. 3), population modeling (fig. 4), ecosystem 

experiments and field surveys (figs. 5, 6, and 7; 

table 13), and comparisons of the sensitivities of 

threatened or endangered species to comparable 

criteria concentrations (table 14).

Criteria Derivation Results

Data Compilation and Evaluation

Stephan and others (1985) provided general guidelines 

on data acceptability for use in criteria calculations. Following 

the review of a large number of relevant studies, more specific 

considerations about which data were useful were made as 

follows. 

The guideline not to use studies “if too many organisms 

in the control treatments died” requires judgments of how 

many deaths are too many. In some instances, specific criteria 

have been published. For example, in acute testing of fishes, 

macroinvertebrates and amphibians with exposures of 48–96 

hours, control survival should be at least 90 percent. In 

early life stage testing of fishes, tests should be considered 

unacceptable if following thinning of the embryos, control 

survival is less than 70 percent (American Society for Testing 

and Materials, 1997, 1998). However, in some tests no clear 

threshold was obvious, such as with chronic tests using less 

commonly tested species. In these cases, it was assumed that 

if a concentration-effect was observed with cadmium and if 

controls and low treatment concentrations were statistically 

distinguished from higher treatments by analysis of variance 

(ANOVA), then the observed responses likely were reliable.

Some tests reported concentrations of cadmium or other 

metals (for example, copper, lead, or zinc) at background 

concentrations that approached or exceeded contemporary 

chronic aquatic-life criteria (for example, Pickering and Gast, 

1972; Chapman, 1978b; Rombough and Garside, 1982). 

Because these tests results showed a concentration—response 

to the cadmium treatments and little control mortality, the 

presence of other metals probably were toxicologically 

insignificant or background metals concentrations possibly 

were biased high, artifacts of “pre-clean” field and laboratory 

procedures (Stephan and others, 1994a). These tests, and 

others with no reported background metals concentrations, but 

that were done at the same laboratories were not considered to 

be disqualified as mixture tests or as having pre-exposed the 

test organisms to high concentrations of the test substance or 

other contaminants according to guidelines by Stephan and 

others (1985).

Static, Renewal, and Flow-Through Exposures

Exposures of test organisms to test solutions are usually 

done through variations on three techniques. In “static” 

exposures, test solutions and organisms are placed in chambers 

and kept there for the duration of the test. The “renewal” 

technique is like the static technique except that test organisms 

are periodically exposed to fresh test solution of the same 

composition, usually once every 24 or 48 hours, by replacing 

nearly all the test solution. In the “flow-through” technique, 

test solution flows through the test chamber on a once-through 

basis throughout the test (American Society for Testing and 

Materials, 1997). The term “flow-through test” is sometimes 

misunderstood to be a test with flowing water that mimics a 

lotic environment. Rather, the term refers to the once-through, 

continuous delivery of test solutions (or frequent delivery 

in designs using a metering system that cycles every few 

minutes). Flows on the order of about 5-volume replacements 

per 24 hours are insufficient to cause appreciable flow 

velocities.

Stephan and others (1985, p. 22) advise that for “some 

highly volatile, hydrolysable, or degradable materials it is 

probably appropriate to use only results of flow-through tests 

in which the concentrations of the material were measured 

often enough using acceptable analytical methods.” Because 

at environmentally realistic concentrations, cadmium is 

highly soluble and is not highly volatile, hydrolysable, or 

degradable, this aspect of the guidelines would not appear to 

be relevant in this analysis. However, following this broad 

statement of principle, in guidelines for calculating species 

mean acute values (SMAVs) preference is given to flowthrough tests when available, but if data from flow-through 

tests are not available for a species, then renewal tests where 

only the initial concentrations were measured, followed by 

data from static exposures in which the actual (as opposed 

to intended) concentrations were never measured. Renewal 

tests that measured test concentrations at the beginning and 

ending of the test were not mentioned (Stephan and others, 

1985, p. 29-30, 36). American Society for Testing and 

Materials (1997) similarly gives a preference for renewal or 

flow-through tests over static tests, 

12 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
“...because the pH and concentrations of dissolved 

oxygen and test material are maintained at desired 

levels and degradation and metabolic products 

are removed, tests using renewal and flow-through 

methods are preferable and may last longer than 96 

h; test organisms may be fed during renewal and 

flow-through tests. Although renewal tests might be 

more cost-effective, flow-through tests are generally 

preferable.”

In the 1985 guidelines, the rationale for the general 

preference for flow-through exposures was not detailed, but 

probably was based on assumptions that static exposures will 

result in LC50s that are biased higher (apparently less toxic) 

than comparable flow-through tests (for example, Pickering 

and Gast, 1972). Additionally, flow-through tests are assumed 

to have more stable exposure chemistries and will result in 

more precise LC50 estimates. When testing with cadmium, 

the assumption of test bias was not supported by recent 

information.

Test bias—Pickering’s and Gast’s (1972) study of acute 

and chronic responses of fathead minnows to cadmium may 

have been one basis for the recommendation that flow-through 

bioassays be used preferentially over static bioassays. The 

flow-through LC50s were lower than the static LC50s (~4,500–

11,000 μg/L for flow-through tests versus about 30,000 μg/L 

for static tests). The fish used in the static tests were described 

as “immature” weighing about 2 g (2,000 mg). The size of the 

fish used in Pickering and Gast’s (1972) flow-through acute 

tests was not reported, but probably was similar. In contrast, 

8- to 9-day old fathead minnow fry usually weigh about 

1 mg or less (U.S. Environmental Protection Agency, 2002c). 

Using newly hatched fry weighing about 1/1,000th of the fish 

used in the 1960s by Pickering and Gast (1972), more recent 

cadmium LC50s for fathead minnows at similar hardnesses 

tend to be around 50 μg/L with no obvious bias for test 

exposure (table 16). The use of large fish probably contributed 

to the exorbitantly high LC50s obtained, and at very high 

concentrations of about 30,000 μg/L may have exceeded 

solubility limits resulting in undetected (if not measured) 

losses from solution during the test. This indicated that fish 

actually experienced lower concentrations over the course of 

the test than measurements at the beginning of the test would 

indicate, resulting in LC50 estimates that were biased high.

More recent testing with copper toxicity to fathead 

minnows has shown that for some test combinations, flowthrough tests tended to produce lower LC50s than static or 

renewal tests (Santore and others, 2001). However in contrast 

to earlier recommendations favoring flow-through testing, 

Santore and others (2001) suggested that flow-through tests 

were biased low because copper complexation with organic 

carbon, which reduces acute toxicity, is not instantaneous and 

typical flow-through exposure systems allowed insufficient 

hydraulic residence time for complete copper-organic carbon 

complexation to occur. (Similar findings were made in tests 

with cadmium and carbonate complexation; Davies and 

Brinkman [1994b]). 

Consequently, it could be argued that flow-through tests 

with low hydraulic residence times could be less relevant to 

field conditions and should not be used if renewal or static 

tests are available. However in this report, flow-through and 

renewal tests were used with equal priority in regressions 

and to estimate species mean acute values. This also is the 

approach used in U.S. Environmental Protection Agency’s 

(2003a) update of aquatic-life criteria for copper. Because 

hydraulic residence time and contact time in field settings 

would be variable, rather than assuming that one type of 

exposure technique best represents field conditions, if 

static, renewal, and flow-through tests otherwise met data 

acceptability guidelines, all were used.

The decision in this analysis not to give singular 

importance to the exposure technique is further supported 

by examples of cadmium acute toxicity. For rainbow trout, 

a well-tested and sensitive species, the two most sensitive 

values (<1 μg/L when adjusted to a hardness of 50 mg/L) 

were obtained using flow-through and renewal testing 

(table 15). However, the two least sensitive acute rainbow 

trout values also were obtained with flow-through exposures 

(>9 μg/L when adjusted to a hardness of 50 mg/L). For 

fathead minnows, in contrast with flow-through values of 

>4,500 μg/L at a hardness of 200 mg/L obtained by Pickering 

and Gast (1972), static values of about 1,000 times lower were 

obtained in tests in even harder water (280 mg/L), but using 

newly hatched fish that were only 24-48 hours old (table 15). 

Phipps and Holcombe (1985) obtained an LC50 of 1,500 μg/L 

in flow-through testing using Lake Superior water with a 

hardness of 44 mg/L; this value is about 100 times higher than 

acute values obtained through static testing with young fry in 

Lake Superior water at similar hardness (table 15). Pickering 

and Gast (1972) and Phipps and Holcombe (1985) both used 

about the same size fish, 0.6 g. Similar results were reported 

with brook trout. One each flow-through and static acute tests 

with brook trout were located, both conducted in waters of 

similar hardness (41–47 mg/L). The LC50 of the static test 

that used fry was less than 1.5 μg/L where the LC50 of the 

flow-through test using yearlings was greater than 5,000 μg/L 

(table 17). These examples show that factors other than test 

type (flow-through or static), especially the life stage of 

exposures, can have more influence on acute toxicity of fish to 

cadmium. Static, renewal, and flow-through test values were 

all used unless inspection of individual test results suggested 

bias.

Tests with No Analytical Confirmation of Exposure 

Concentrations

In aquatic toxicity testing, treatment concentrations 

often are described as “nominal” for intended target test 

concentrations and “measured” for concentrations measured 

through laboratory analysis of the chemical in water. Stephan 

and others (1985) recommended using the results of tests with 

measured concentrations in preference to those for which the 

intended nominal concentrations were not confirmed. 

Results 13
However, they also recommended that 

“...for a species for which no such result is 

available, the species mean acute value should be 

calculated as the geometric mean of all available 

acute values, i.e., results of flow-through tests 

in which the concentrations were not measured 

and results of static and renewal tests based on 

initial concentrations (nominal concentrations 

are acceptable for most materials if measured 

concentrations are not available) of the test 

material.” 

The limited acceptance of some unmeasured test 

values probably was partly influenced by the practical 

consideration that minimum species diversity requirements 

might not be met with available data for some chemicals if 

all unmeasured results were excluded. Sufficient data were 

available to derive criteria using only test values based on 

measured concentrations for the cadmium dataset compiled 

here. Further, for the acute and chronic datasets, some of the 

lowest reported results were for species for which the only 

test values located were from unmeasured exposures. These 

species (striped bass, Morone saxtilis and the cladoceran 

Moina macrocopa) also were the only representatives of 

their genus. In the calculations of the final acute and chronic 

criteria values, the acute and chronic datasets were censored 

so that only the four most sensitive genus values were used. 

Thus, in preliminary calculations the final acute and chronic 

values were influenced by questionable values from single test 

results, because the intended exposure concentrations were 

not measured. It therefore seemed appropriate to exclude from 

the final acute and chronic value calculations any influential 

values that were questionable because they were obtained 

from unmeasured exposures. However, instead of selectively 

excluding only unmeasured values that appeared sensitive, it 

was more straightforward to simply exclude all test values that 

were obtained without analytical confirmations of exposure 

concentrations. Several noteworthy tests that used unmeasured 

exposures are listed in table 17, such as tests with unusually 

low effects values or tests with threatened or endangered fish 

or molluscs, or closely related surrogates.

Life-Cycle Versus Shorter-Term Chronic Data

For this analysis, data from long-term “chronic” 

exposures were used quantitatively (for example, in hardnesstoxicity regressions and in species mean value calculations) if 

they were obtained using life-cycle methods or from shorter 

term methods for estimating chronic toxicity if the results 

obtained with the shorter-term methods were more sensitive, 

or similar in sensitivity to those results obtained from life 

cycle testing. This is a modification from Stephan and others 

(1985) recommendations to use life-cycle or partial life-cycle 

results in preference to shorter-term methods when data 

from both are available. This modification has the practical 

advantage of increasing the amount of available data, such 

as allowing the development of hardness—chronic toxicity 

relations in lieu of applying hardness—acute toxicity relations 

to chronic data. For fish species, life-cycle data are too limited 

to develop such relations. More importantly, in the 20 years 

since Stephan and others (1985) made their recommendations, 

more data have become available that indicate that, at least for 

fish, the life stage at which long-term exposures are initiated 

(for example, embryo or newly hatched) can influence the test 

sensitivity at least as much as the test duration. 

Shorter-term methods that test apparently sensitive 

life stages were favorably compared with longer-term life 

cycle or partial life cycle tests, based on similar sensitivities 

of responses (McKim, 1977, 1985; Norberg-King, 1989). 

However, Suter and others (1987) cautioned that earlylife stage testing sometimes can underestimate the toxicity 

of substances as compared in life-cycle tests. Regardless, 

numerous tests have shown that reduced sensitivity of either 

life cycle or early-life stage tests may occur if the tests are 

initiated during an insensitive life stage (for example, embryos 

or adults). Initiating metal exposure at the resistant embryo 

stage may allow acclimation, which may mute later responses 

of otherwise sensitive life stages (Sinley and others, 1974; 

Spehar, 1976; Chapman, 1978a, 1982, 1994; Spehar and 

others, 1978b; Brinkman and Hansen, 2004, 2007). 

Brinkman and Hansen (2007) conducted three pairs of 

chronic tests with brown trout in which exposures began at 

the embryo stage (early-life stage test) or at the swim-up fry 

stage (juvenile growth and survival test). Tests initiated at 

the swim-up fry stage and with a 30-day duration were 2–3 

times more sensitive than paired tests initiated at the embryo 

stage and maintained for 55 days. In contrast, Norberg-King 

(1989) compared the effects of zinc exposures to newly 

hatched fathead minnows that either were pre-exposed to 

zinc as embryos or not and found little difference from the 

pre-exposures. Results of comparative testing with copper 

vary. Chapman (1994) found that the most sensitive test for 

rainbow trout was exposure beginning at swim-up with no 

prior exposure to provide acclimation. The least sensitive test 

was to begin exposure either prior to hatch at the eyed stage 

or at hatch. Initiating exposures at these stages provided some 

opportunity for acclimation prior to reaching the sensitive 

swim-up stage. However, Besser and others (2005) observed 

little difference in results of copper toxicity tests with rainbow 

trout as swim-up fry and maintaining exposures for 30 days 

or by initiating tests as embryos in classic early-life stage 

tests and maintaining exposures for 60 days. In summary, 

the available literature with metals indicate that for shorterterm (shorter than life cycle) methods for estimating chronic 

toxicity, juvenile growth and survival tests initiated as fry 

often, but not always, provide more sensitive data than earlylife stage initiated as embryos. Thus in this report, chronic 

values from each type of test were used without giving any 

priority to one over the other. Of the data reviewed, no lifecycle test results with cadmium and fish were appreciably 

more sensitive than early-life stage test results with the same 

14 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
species. Therefore, juvenile growth and survival, early-life 

stage tests, and life-cycle tests all were used together to 

estimate mean chronic values for fish species.

Limiting estimates of chronic values to life-cycle test 

data would have the practical disadvantage of limiting the 

chronic dataset for fish to a small, static dataset that mostly 

was generated 20–40 years before this report. This is because 

for most fish species, life cycle tests of about 9 months to 

more than 3 years are extremely labor intensive, costly, and 

have been infrequently published. For example, no studies of 

cadmium and fish with life-cycle toxicity tests were reviewed 

that were conducted within the last 10 years, two within 

the last 20 years, and about 6 more within the last 40 years 

(Pickering and Gast, 1972; Benoit and others, 1976; Brown 

and others, 1994; U.S. Environmental Protection Agency, 

2001). 

The fathead minnow is the standard fish species used 

to estimate chronic effects of effluents in regulatory toxicity 

tests (U.S. Environmental Protection Agency, 2002c). Results 

of two life-cycle tests with cadmium conducted by Pickering 

and Gast (1972) in the mid 1960s were about 2–5 times higher 

than results of recent 7-day short-term chronic tests that were 

conducted at higher hardness levels (table 17). The life-cycle 

tests were initiated with much larger (about 600 mg) fish 

than those used in the contemporary short-term tests, which 

commonly weigh <1 mg at the end of the test (for example, 

Castillo and Longley, 2001; U.S. Environmental Protection 

Agency, 2002c).

Fewer comparative data on exposure duration and 

responses to cadmium were located for invertebrates than 

for fish. Winner (1988) found that a 7-day exposure was 

insufficient to elicit sensitive responses from Daphnia magna

in comparisons with 7-day Ceriodaphnia exposures or 21-day 

Daphnia magna exposures. With the amphipod Hyalella 

azteca, Ingersoll and Kemble (2001) obtained LC25 values of 

2.7 to 2.1 μg/L from 10- to 28-day exposures, respectively, 

with little additional decrease from 28 to 42 days. In contrast, 

in regression analyses of responses of Hyalella azteca to 

cadmium over a hardness range of 17–280 mg/L, most 

variability in responses was accounted for by hardness, despite 

exposure durations varying from 14 to 42 days. 

An issue related to test exposure duration was whether 

exposures that were interrupted shortly before planned test 

termination could still provide useful data. Exposure durations 

in early-life stage tests with salmonids should be at least 60 

days (American Society for Testing and Materials, 1998). 

However, Brinkman and Hansen (2007) reported results from 

three tests with brown trout that were exposed for only 55 

days and Mebane and others, 2008) obtained results from a 

rainbow trout test that exposed fish for only 53 days. Similarly, 

Chadwick Ecological Consultants, Inc. (2004a) exposed 

Daphnia pulex to cadmium for only 18 days instead of the 

21 day protocol duration. For these situations, the data were 

included in the analyses following a case-by-case review that 

compared the hardness-adjusted results to those obtained from 

other tests with the same species, and where data on responses 

over time were reported, whether or not mortality rates had 

leveled off. 

Statistical Interpretation of Chronic Tests

Chronic data particularly were evaluated carefully 

because previous investigators used a variety of terms and 

definitions to report results, and because no consistent 

guideline on what magnitude over effect is considered an 

“unacceptable” effect. Stephan and others (1985) advise 

that a chronic value may be obtained by calculating the 

geometric mean of the lower and upper chronic limits from 

a chronic test or by analyzing chronic data using regression 

analysis. Lower and upper chronic limits often were based on 

statistical hypothesis testing and correspond to the no- and 

lowest-observed-effect concentrations. In aquatic toxicology 

literature, typically, the no-observed-effect concentration 

(NOEC) is the highest test concentration that results in 

responses that are not statistically different from the control 

responses. The lowest-observed-effect concentration (LOEC) 

is the first treatment that is greater than the NOEC. Between 

these “chronic limits,” is a hypothetical maximum acceptable 

toxicant concentration (MATC), which is an assumed 

threshold for toxic effects. The point estimate of a MATC is 

the geometric mean of the NOEC and the LOEC (McKim, 

1985).

Stephan and others (1985) used precise language to 

describe how chronic values may be obtained with a subtle 

but important difference from the usual NOEC and LOEC 

definitions. The term “not statistically different” was replaced 

with “did not cause an unacceptable amount of adverse 

effect1” (The full definition of chronic limits is given in the 

glossary.) This language recognized important limitations 

inherent in statistical hypothesis testing for “significance.” 

Chronic responses often are considered significant if statistical 

hypothesis testing indicates less than a 5 percent likelihood 

(a significance level of 0.05) that apparent differences in test 

responses are due simply to chance. However, in hypothesis 

testing, biologically important adverse effects may not be 

statistically significant due to high variability, small sample 

sizes, or other statistical vagaries. Biologically trivial effects 

might be statistically significant if variation is low or sample 

1

 The complete definitions of chronic limits were, “A chronic value may 

be obtained by calculating the geometric mean of the lower and upper chronic 

limits from a chronic test or by analyzing chronic data using regression 

analysis. A lower chronic limit is the highest tested concentration (a) in an 

acceptable chronic test, (b) which did not cause an unacceptable amount of 

adverse effect on any specified biological measurements, and (c) below which 

no tested concentration caused an unacceptable effect. An upper chronic 

limit is the lowest tested concentration (a) in an acceptable chronic test, (b) 

which did cause an unacceptable amount of adverse effect on any specified 

biological measurements, and (c) above which all tested concentration also 

caused such an effect.” (Stephan and others 1985, p. 39)

Results 15
sizes are large. Because chronic toxicity tests commonly use 

about two to five sample replicates, the problem of important 

adverse effects not being statistically significant probably is 

the more likely of the two outcomes in most tests. Cases of 

the statistical MATCs and NOECs corresponding to fairly 

high levels of adverse effect have been documented (Suter and 

others, 1987; Crane and Newman, 2000). Stephan and others 

(1985) were aware of these limitations and advised that 

“...the amount of effect that is considered 

unacceptable is often based on a statistical 

hypothesis test, but might also be defined in terms 

of a specified percent reduction from the controls. 

A small percent reduction (e.g., 3 percent) might 

be considered acceptable even if it is statistically 

significantly different from the control, whereas a 

large percent reduction (e.g., 30 percent) might be 

considered unacceptable even if it is not statistically 

significant).” 

However, beyond this example, Stephan and others (1985) 

made no further recommendations for specifying what percent 

reductions from the controls would be considered acceptable 

or unacceptable. Few recommendations from the refereed 

literature were helpful to define what level of effect x for the 

effects concentration adversely affecting p percentage of a test 

population (ECp) is “biologically acceptable” for biological 

communities. Published views referring to a level of effect p

for the ECp considered biologically acceptable ranged from 

5 to 25 percent, and were based on professional judgments 

(Bruce and Versteeg, 1992; de Bruin and Hof, 1997; van der 

Hoeven and others, 1997).

For this report, chronic values were obtained from 

reported chronic tests as follows. Generally, chronic values 

were calculated as MATCs, the geometric mean of the NOECs 

and LOECs. If the percentage of reductions in survival, 

growth, reproduction, etc., associated with the NOECs and 

LOECs were readily apparent in the source, reductions 

much greater than about 10 percent in tests with fish could 

be “unacceptable” and reductions much greater than about 

20 percent in tests with invertebrates could be “unacceptable.” 

If MATCs were greatly different from these ECp percent 

reductions, then a test statistic was used that was closer to 

the respective EC10 or EC20 value. For some influential tests, 

ECx values were calculated from the original data if sufficient 

data were presented. In many instances, insufficient data were 

presented to calculate ECp values. 

For fish, this choice of 10 percent as the estimate of 

the upper bound of an “acceptable effect” was supported by 

modeled extrapolations of early-life stage mortality rates to 

wild fish populations. In several scenarios when reductions 

occurred in fish populations that were reasonably stable, 

habitats were intact, and if environmental conditions were 

not otherwise severe, reductions of about 20 percent (EC20) 

in growth or first year survival likely would be sustainable. 

However, in more vulnerable populations, or in populations 

subject to other stressors, reductions of 10 percent (EC10) 

or less for growth or mortality endpoints would be a better 

estimate of an acceptable low-effects threshold (Barnthouse 

and others, 1990; Boreman, 1997; Paul and others, 2003; 

Spromberg and Meador, 2005). For invertebrates, the 

assumption that endpoint reductions much greater than 

20 percent could have “unacceptable” adverse effects on 

populations follows modeled extrapolations of mortality rates. 

For example, Kuhn and others (2002) predicted amphipod 

population growth rates will begin to decline dramatically 

when acute mortality (measured in 10-day tests) starts to 

increase above 20 percent. Concentrations causing invertebrate 

acute mortality rates on the order of 50 percent have resulted 

in extinctions in longer term mesocosm tests (Stark, 2005). 

These examples suggest that reductions much more than 

about 20 percent would not be sustainable in at least some 

invertebrate populations. Although these selected examples are 

not intended to represent the rich literature on extrapolating 

effects from the individual to population, they do give some 

support to the assumptions that reductions on the order of 

10 or 20 percent for fish and invertebrates, respectively, 

approximate a divide between effects that are likely acceptable 

and unacceptable.

The assumption that many invertebrate populations could 

tolerate greater percentage of reductions than fish populations 

is supported by shorter recovery times from disturbances 

for invertebrates compared to fish. For a given toxicological 

sensitivity to a contaminant at a given level of effect, species 

that are less able to recover from disturbance are more 

vulnerable and are at increased risk of extinction. Aquatic 

insect populations conceptually would be less vulnerable 

to contaminant-induced mortality than more long-lived fish 

populations. Aquatic insects commonly have shorter life 

cycles than fish, high dispersal abilities, and generally high 

reproductive potential. Long-lived large, invertebrates such 

as some molluscs are exceptions to these generalizations. 

Empirical and predicted comparisons of recoveries from 

disturbances for many fish and invertebrate populations 

or assemblages found that invertebrates usually recover 

from disturbances faster than fish (Niemi and others, 1990; 

Detenbeck and others, 1992; Barnthouse, 2004).

Of the data reviewed for this report, ECp values were 

calculated and compared with statistical no- and lowestobserved effect concentrations for 19 tests with invertebrates 

and 15 tests with fish. These data showed reasonably good 

agreement between the MATC values and the EC20 and 

EC10 values for invertebrates and fish, respectively. Of the 

invertebrate values, the average MATCs and NOECs were 

greater than corresponding EC20 concentrations (28 and 16 

percent greater, respectively). With fish the average NOEC 

values were 11 percent lower than the corresponding EC10

values, and average MATCs were 16 percent greater than 

EC10 values. This comparison suggests that for cadmium, the 

typical expected adverse effect level associated with a MATC 

probably is about 20–30 percent for invertebrates, and about 

10–15 percent for fish. 

16 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Diversity of Data

About 279 acute values were located in datasets, 

representing 69 species, 58 genera, and 33 families. About 

93 chronic values were located in datasets, representing 

28 species, 21 genera, and 17 families. Although these data 

are limited in diversity compared to the diversity of aquatic 

life in North America (>10,000 animal species, table 2), 

these datasets are more diverse than those available for many 

other chemicals. For example, Forbes and Calow (2002) 

reviewed published chronic SSD-based datasets consisting 

of 5–25 species, and 12–64 species for acute datasets. The 

minimum diversity guidelines for the chronic dataset are 

exceeded for fish and planktonic crustaceans, but are only 

just met for other major taxonomic groups such as insects and 

molluscs (table 2). If in fact, the chronic dataset represented a 

random sample from all fish and invertebrate species available, 

then about 50 percent of the test values would be for insects, 

90 percent for all invertebrates, and less than 10 percent for 

fish. Instead, only 4 percent of the chronic values were for 

insects, 32 percent for all invertebrates, and 60 percent were 

for fish. Among the fish, the richest data were for salmonids, 

which probably is because of their social importance and 

their sensitivity. The larger acute dataset meets or exceeds the 

minimum diversity guidelines for all groups (not shown). 

Some variations in taxonomic nomenclature and spelling 

of common and scientific names of organisms were found in 

the literature review. For this report, the Integrated Taxonomic 

Information System (2002) was used as the taxonomic 

authority for scientific and common names of taxa other 

than fishes. For fishes, Nelson and others (2004) was used 

except for Oncorhynchus mykiss. The common name for 

Oncorhynchus mykiss used in this report is “rainbow trout” 

regardless if the original study referred to specific forms such 

as redband trout, golden trout, or the sea-run steelhead. 

Acute Toxicity of Cadmium to Freshwater 

Animals 

Acute values for Daphnids, for example, Daphnia sp. 

and Ceriodaphnia sp., were based on 48-hour LC50s, all other 

acute values were 96-hour LC50s. Using pooled hardness-acute 

toxicity relations, acute values were adjusted to a hardness of 

50 mg/L to make the responses more comparable. Estimated 

species mean acute values (SMAVs) were calculated as the 

geometric means of the hardness adjusted individual test 

values for a species and the estimated genus mean acute values 

(GMAVs) were calculated from the geometric means of the 

SMAVs (table 15). 

The extreme variability of Daphnia magna acute values, 

which even after normalization vary by more than 100 times, 

is particularly noteworthy. Even among replicated tests in the 

same experiment acute Daphnia values varied by more than 

a factor of 8 (Nebeker and others, 1986a). In cases where the 

available data for a species vary by greater than 10 times, 

Stephan and others (1985) recommend not using some or all 

values in species mean or genus mean values. However, all 

data were used based on the rationale that this variability, 

which is primarily genetic, in part reflects that occurring in 

field populations. Different genotypes or clones of this species 

have been shown to have high variability in sensitivity. This 

variability is assumed to be both less than but at least partly 

reflect that found in field populations (Baird and others, 1990, 

1991; Baird, 1992; Forbes and Depledge, 1992, 1993; Forbes 

and Forbes, 1993b; Barata and others, 1998; Goulden, 1999;

Barata and others, 2000; Barata and others, 2002a). Barata 

and others (2002a) concluded that laboratory selection favors 

individuals with high fitness or reproductive performance 

under optimal laboratory conditions resulting in laboratory 

populations with similar or lower tolerance to toxic stress 

than their original field populations. Given that populations 

can exhibit high levels of genetic variability in tolerance to 

toxic stress, minimizing genetic diversity in toxicity tests will 

increase the uncertainty attendant in extrapolating from the lab 

to the field. 

Overall, available data indicate that salmonids, sculpins, a 

darter, and the amphipod Hyalella are distinctly more acutely 

sensitive to cadmium than other taxa. When rank-ordered by 

genus mean acute values, the four most sensitive genera were 

Oncorhynchus, Salvelinus, Salmo, and Cottus (table 7).

Chronic Toxicity of Cadmium to Freshwater 

Animals

The most consistently sensitive species was the amphipod 

Hyalella azteca with seven hardness-adjusted values ranging 

from 0.14 to 0.51 μg/L. The second most sensitive genus 

value was for a sculpin, Cottus, followed by an amphipod, 

Gammarus fasciatus, and the trout Salvelinus (table 7).

As with acute data, Daphnia magna values were 

noteworthy for their variability, varying by over a factor of 20. 

As with the acute data, this variability probably at least partly 

reflects that found in field populations. Therefore, no values 

were discarded solely based on variability. Methodological or 

test condition differences likely contributed to the variability 

in the results. For example, the three lowest chronic values 

for Daphnia magna (0.13–0.18 μg/L, hardness adjusted) were 

from a study by Chapman and others (1980). Winner (1986) 

suggested these results may have been biased low because 

hardness was adjusted by diluting test-water with deionized 

water and deionized waters may contain chemicals leached 

from the deionized resin, which was very toxic to Daphnids. 

Diet may have played some role because the animals used 

by Chapman and others (1980) were fed a trout-chow diet 

and Daphnia magna fed a trout-chow diet were much more 

sensitive to copper than those fed a vitamin enriched algal diet 

(Winner, 1986).

Results 17
Presumably, other Daphnids or other species also might 

have similar genetic variability, but no similar research to that 

reported for Daphnia magna was reviewed. Thus, a chronic 

value obtained with Ceriodaphnia dubia by Jop and others 

(1995), that was 31 times higher than the lowest hardness 

adjusted value for this species was not used, even though 

the test appeared otherwise acceptable. The remaining seven 

chronic values for the genus Ceriodaphnia were within 7 times 

of each other. Within the genus Salvelinus, the value for lake 

trout was greater than 10 times the lowest value for the genus, 

which was obtained with bull trout. The lake trout value was 

not used in the GMCV calculations. The remaining five values 

for Salvelinus were all within 3 times of each other (table 16). 

In the case of fathead minnow, Pimephales promelas, data, 

the highest hardness-adjusted chronic value, generated from 

a life-cycle test, was 12 times higher than the lowest value, 

which was generated from a 7-day growth and survival test. 

In this case, these life-cycle tests were retained because of the 

preference for life-cycle test data in estimating species mean 

chronic values (Stephan and others, 1985, p. 38), despite the 

risk that these results possibly were biased high. 

Few, if any, of the chronic datasets reviewed were ideal. 

Results published as journal articles were sometimes highly 

summarized, omitting details such as what magnitude of 

responses were considered no- and lowest-observed effect 

concentrations (for example, Eaton, 1978). In what were 

likely tradeoffs between replication, numbers of treatments, 

and test durations, some tests were unreplicated or only used 

two replicates, which were insufficient for statistical testing 

for differences between treatments using ANOVA (Benoit 

and others, 1976; Sauter and others, 1976; Davies and others, 

1993).

Other Data on the Effects of Cadmium to 

Freshwater Animals.

“Other data” that for various reasons were not directly 

used to develop criteria, but were still pertinent were 

evaluated and compared to the acute and chronic criteria 

values (table 17). Stephan and others (1985) considered test 

endpoints relating to mortality, growth, or reproduction to be 

biologically important and they specifically called for using at 

least these endpoints in criteria derivation. Although nothing 

was noted in their guidelines to preclude including other 

biological responses in criteria derivation, in practice criteria 

documents have focused on these endpoints. However, other 

pertinent data are available on adverse effects of cadmium 

to aquatic organisms. Numerous reports were reviewed that 

evaluated effects of cadmium on freshwater animals including 

acclimation, biochemical effects, swimming performance, 

predator-prey interactions, bioaccumulation and elimination, 

and factors affecting toxicity. Such data could, however, affect 

a criterion if test concentrations were measured, the endpoint 

was biologically important, and if the data were obtained with 

a biologically important species (Stephan and others, 1985).

Tests compiled in table 17 were considered pertinent 

for estimating risks or response thresholds for freshwater 

organisms to cadmium, but for some reasons it did not seem 

appropriate to pool with the acute and chronic data listed in 

tables 15 and 16. Reasons for treating these data separately 

included: 

1. Tests with durations or test statistics that could not 

be directly compared with other data (for example 

acute exposures longer than 96-hours, incipient 

lethal levels reported in lieu of LC50s); 

2. Effects were observed in potentially important 

biological endpoints, such as swimming impairment 

or reduced prey capture efficiency in salmonids, but 

response thresholds could not be estimated from 

the concentrations tested. In these cases unbounded 

LOECs were obtained, that is, effects were 

observed at the lowest treatment tested. Often, it 

did not seem appropriate to interpolate between the 

unbounded LOEC and the near-zero control NOEC 

concentration; 

3. Tests of biochemical or physiological change that 

were not accompanied by organism level effects 

such as reduced survival or growth;

4. Tests of acclimation influences on subsequent 

toxicity or tests that otherwise appeared to have 

substantial pre-exposure to cadmium;

5. Tests that used unusual dilution waters, for which the 

general assumption that hardness is correlated with 

calcium concentrations, alkalinity, and pH would not 

hold;

6. Tests that used insensitive life stages; some values 

obtained using more sensitive life stages are shown 

here for contrast but also were included in table 15;

7. Data in a publication which were identical to data 

reviewed in another source were assumed to be the 

same data described in more than one publication 

and were used only once in tables 15 or 16;

8. Tests with questionable results, such as tests with 

unmeasured exposure concentrations, or that 

reported insufficient information to evaluate.

Some of the “other” tests summarized in table 17

indicated adverse effect concentrations lower than calculated 

criteria values. In unmeasured exposures, an estimated chronic 

value of 0.28 μg/L was obtained for the cladoceran Moina 

18 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
macrocopa, which was lower than the calculated chronic 

criterion of 0.52 μg/L at a hardness of 82 mg/L (table 17). 

This value is slightly higher than the lowest values used in the 

chronic dataset for a cladoceran (Daphnia magna, table 16). 

The foraging success of lake trout, Salvelinus namaycush on 

fingerling rainbow trout, Oncorhynchus mykiss, was reduced 

after both species were exposed to 0.5 μg/L cadmium for 

8–9 months. This reduced foraging success suggests a low 

response threshold for cadmium-caused behavioral changes 

(table 17). Whether that response would be expected in more 

natural conditions is unclear because a population of lake 

trout that was exposed to cadmium through a whole-lake 

experiment did not show reduced size or condition factor 

compared to baseline or reference conditions (this report 

“Cadmium Whole-Lake Ecosystem Experiment” section). 

Another test reporting adverse effects below the calculated 

chronic criterion involved reduced hatching success of 

rainbow trout embryos after their parents were exposed to 

0.2 μg/L of cadmium for 18 months (W.J. Birge and J.A. 

Black, unpub. data, as cited by Birge and others, 1981). 

However, because the information reported for this test was 

very limited, the results were difficult to interpret and thus the 

significance of this test is unclear (table 17).

An effort was made to locate data on effects of cadmium 

on snails because several snail species are considered 

endangered in Idaho and because of concerns that pulmonate 

(air breathing) aquatic snails are sensitive to copper and lead, 

and were fairly sensitive to zinc (Grosell and Brix, 2004; 

Grosell and others, 2006). Several studies of the responses 

of the common pond snail, Lymnaea spp., to cadmium were 

located, including the two chronic studies listed in table 17. 

The LOECs were 20–40 times higher than chronic criterion 

values for the test hardnesses, but these data are ambiguous 

because adverse effects were observed at the lowest 

concentrations tested. 

Hardness–Toxicity Relations

Because previous investigations have found that a log-log 

relation fits data for acute toxicity of metals and hardness well 

(Stephan and others, 1985; Meyer, 1999), logarithms were 

used for evaluating hardness-toxicity relations for cadmium 

(fig. 2; tables 3 and 4). 

Acute values obtained over a broad enough range of 

hardness values to be useful to evaluate apparent hardnesstoxicity relations were available for 11 species over hardnesses 

ranging from 5 to 290 mg/L (fig. 2). The slopes of hardnessacute toxicity relations ranged from 0.4 to 1.5, with this 

relation explaining from 32 to 99 percent of the variability 

in species acute values When the slopes were pooled using 

species as a grouping variable (Zar, 1984), the overall 

hardness-acute toxicity slope for all species was 0.8368 (table 

3).

Chronic values obtained over a broad enough range of 

hardness values to be useful to evaluate apparent hardness–

toxicity relations were available for seven species. Data were 

available for hardnesses ranging from 17 to 280 mg/L (fig. 2). 

In contrast to the acute slopes, the slopes of hardness–chronic 

toxicity relations were remarkably similar between the species, 

and when plotted were nearly parallel (fig. 2). Individual 

regression slopes ranged only from 0.50 to 0.77, and for 5 

of the 7 species, this simple log-linear least-squares relation 

explained greater than 90 percent of the variability in the data. 

When the slopes were pooled using species as a grouping 

variable (Zar, 1984), an overall hardness–chronic toxicity 

slope for all species of 0.6247 was obtained (table 4).

It is noteworthy that this hardness-toxicity relation 

accounted for greater than 90 percent of the variability in 

chronic test results because the data with brook trout, brown 

trout, rainbow trout, and Hyalella resulted from varying 

“chronic” exposure durations. Durations of the chronic tests 

ranged from 60 to 1,100 days for brook trout, 55 to 665 days 

for brown trout, 53 to 665 days for rainbow trout, and 14 to 42 

days for Hyalella azteca (table 16). 

In several cases, the data used in the regressions were 

censored to minimize the influence on confounding variables 

such as sensitivity differences from different life stages and 

possible differences between animal strains or test methods. 

Data were restricted more for inclusion in the hardness–

toxicity analyses than for the overall calculations of species 

mean acute or chronic data. This difference was because to 

evaluate hardness–toxicity relations all variability from other 

factors is undesirable, whereas to evaluate species sensitivity 

rankings, obtaining the most representative data for sensitive 

life stages of a species probably is more important than 

minimizing variability so that some variability from different 

stocks, cultures, or other unknown factors is tolerable. Specific 

test values that were used in the hardness–toxicity regressions 

are indicated in the notes on tables 15 and 16.

For Daphnia magna, only data from Chapman and others 

(1980) were used in both the acute and chronic hardness 

regressions. This was the only study reviewed that explicitly 

tested organisms from the same culture, using the same testing 

facility and investigators, and conducted tests in waters with 

multiple (≥ 3) hardnesses. This censoring resulted in a chronic 

slope estimate for this species of 0.77 with a regression 

coefficient (R2) of 0.96, whereas if all otherwise acceptable 

chronic Daphnia magna data were used, no hardness-toxicity 

relation would result (slope of 0.2, R2 of 0.007, P=0.8 (tables 4

and 5). Similarly, to reduce variability that may have resulted 

from different cultures or laboratory waters, only chronic 

Daphnia pulex data generated at Miami University were used.

The overall slopes obtained for acute and chronic 

toxicity–hardness relations of 0.84 and 0.62 are lower than 

the about 1.0 slope for acute toxicity that Meyer (1999) 

predicted for transition metals. A possible reason for these 

Results 19
Figure 2. Hardness-toxicity relations for acute and chronic data. Data are from studies in tables 15 and 16 with 

an “h” in the code/notes column.

id91UO5_fig02

h

h

h h h h

b b

b

b r

r

r

r r

r

m

m

m

p

B p p p

B

B

B

B

A

A

A

L

L t

t

t

G

G

g

g

g

v

v

m

mm

mm

f

f

f ff

f

B

B

B

B

s

s

s

sss

r

r

r

r rr r

r

r

r

r

r

r

r

r

rr

r

r

r

r

r

r r r

r

h h B

r r r

r

r r

r r

r

h h

r

-1

0

1

2

3

4

5

0.5 1 1.5 2 2.5 3

Logarithms of chronic toicity values

0.5 1 1.5 2 2.5 3 -1

0

1

2

3

4

5

Logarithms of acute toxicity values

Logarithms of hardness values

Logarithms of hardness values

Green sunfish

g Bluegill

L Limnodrilus hoffmeisteri

t Tubifix tubifix

v Southern rainbow

Daphnia magna

f Fathead minnow

h Hyalella azteca

s Bull trout

B Brown trout

r Rainbow trout

Daphnia magna

r Rainbow trout

B Brown trout

Brook trout

h Hyalella azteca

Daphnia pulex

Aelosoma

p

A

m

m

G

b

Acute

Chronic

20 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 3. Acute toxicity values versus hardness for individual and pooled regressions.

[n: number of observations. Slope: Slope of the relations between the natural logarithm (ln) of hardness values and ln toxicity values. R2: Coefficient of 

determination or fraction of the variance explained by the regression (Helsel and Hirsch, 2002). P: probability that the apparent relations between toxicity and 

hardness actually resulted from chance alone and that the actual slope of the linear regression is zero. Underlining emphasizes P-values that are significant at 

0.05 or less. mg/L, milligram per liter; –, missing or inapplicable value; <, less than]

Species n Slope R2 P (significance) 95-percent 

confidence limit

Daphnia magna1 5 1.182 0.91 0.01 0.52, 1.84

Worm, Limnodrilus hoffmeisteri 2 .788 – – None

Worm, Tubifix tubifix 3 .6238 .93 .17 -1.56, 2.81

Southern rainbow (mussel), Villosa vibex 2 .9286 – – None

Amphipod, Hyalella azteca2 4 .7866 .89 .06 -0.07, 1.64

Fathead minnow3 6 .6222 .316 <.01 0.46, 0.78

Brown trout 5 1.193 .973 <.01 0.83, 1.55

Bull trout 6 1.548 .696 .04 1.29, 2.96

Rainbow trout 35 1.036 .573 <.01 0.72, 1.35

Green sunfish 2 .4220 – – None

Bluegill 4 .9111 .99 <.01 0.627, 1.19

Pooled slope for all species 74 0.8403 0.978 <0.01 0.676, 1.00

1 Data from Chapman and others (1980).

2

 Using the mean of six values that were all generated by Collyard and others (1994) at a hardness of 90 mg/L as one value.

3

 Using only the results of acceptable tests with fathead minnow fry as shown in table 16.

Species n Slope R2 P (significance) 95-percent 

confidence limits

Aelosoma headleyi 3 0.7429 0.7855 0.30 -4.19, 5.67

Daphnia magna1 3 .7712 .962 .12 -1.67, 2.71

Daphnia pulex2 6 .5039 .617 .21 0.348, 0.713

Hyalella azteca3 6 .6853 .93 <.01 0.42, 0.94

Brook trout 4 .6187 .980 .01 0.348, 0.889

Brown trout4 5 .6987 .926 <.01 0.335, 1.06

Rainbow trout4 6 .5300 .942 <.01 0.348, 0.712

Pooled slope for all species 31 0.6247 0.991 <0.001 0.533, 0.716

1

Using only data from Chapman and others, (1980).

2

Miami University data only.

3

Excluding Borgmann and others (1991) value as anomalous.

4

Early life stage and life cycle tests only.

Table 4. Chronic toxicity values versus hardness for individual and pooled regressions.

[n: sample size. Slope: slope of the relations between the natural logarithm (ln) of hardness values and ln toxicity values. R2: coefficient of 

determination or fraction of the variance explained by the regression (Helsel and Hirsch, 2002). P, probability that the apparent relations 

between toxicity and hardness actually resulted from chance alone and that the actual slope of the linear regression is zero. Underlining 

emphasizes P-values that are significant at 0.05 or less. <, less than]

Results 21
lower slopes is that with cadmium the mitigating influence of 

increasing hardness on acute cadmium toxicity is offset by the 

corresponding aggravating influence of increasing pH on acute 

cadmium toxicity. 

Extrapolating the calculated hardness–toxicity relations 

much beyond the tested range of 5 to 290 mg/L as calcium 

carbonate would be questionable. Working with Daphnia 

magna and zinc, which has some ecotoxicological similarities 

to cadmium, Heijerick and others (2002) found that at 

hardnesses greater than 325 mg/L as calcium carbonate, no 

linearity, and even a decrease in 48-hour EC50s was observed. 

At the low end of the range, further extrapolation of the 

relations probably is moot because few ambient waters have 

hardness much less than 5 mg/L. At least in the case of 

salmonids, most reported cadmium toxicity during chronic 

tests was actually acute mortality that either occurred in the 

first 4–5 days in tests initiated with fry, or in the first 4–5 days 

after swim-up occurred in tests initiated at the more resistant 

egg stage (McDonald and Wood, 1993; Hollis and others, 

1999, 2000b; Hansen and others, 2002b; Brinkman and 

Hansen, 2007; Mebane and others, 2008). As a result, acute 

to chronic ratios with salmonids tend to be close to 1.0 (table 

6). Therefore, at least for salmonids, it follows that the tested 

hardness–chronic toxicity relations probably would be relevant 

over the range of hardness–acute toxicity relations.

Newman (1991) presented arguments why a statistical 

bias encountered with the backtransformed, least-squares 

regression models used to develop metals criteria can 

compromise the accuracy of associated predictions. In datasets 

selected from metals criteria documents available at the 

time, Newman found the bias was as high as 57 percent and 

suggested that hardness-based metals criteria equations would 

benefit from additional terms to correct for this bias. Using his 

method of estimating this bias, the hardness adjustments for 

acute values were estimated to be low biased by 17 percent, 

and the hardness adjustments for chronic values were 

estimated to be low biased by less than 2 percent. Because 

the potential bias of these hardness-adjusted values seemed 

low, the potential benefit of improving model precision gained 

through adding additional terms to the cadmium criteria 

equations seemed offset by a more complex criteria equation 

that would have a different form than other metals criteria.

Acute to Chronic Ratios

Acute to chronic ratios (ACRs) were calculated speciesby-species to provide a tool for estimating chronic effects 

thresholds from acute data. Species mean ACRs were 

estimated for 17 species using 38 pairs of acute and chronic 

tests (table 6). Only pairs of tests that were conducted using 

the same dilution water source were used. Some test pairs 

were excluded because the acute tests were conducted with 

resistant, older fish and the resulting ACRs calculated from 

these tests would be biased high. 

When species mean ACRs seem to increase as the 

SMAVs increase, Stephan and others (1985) recommended 

that the final ACR be calculated with species that have SMAVs 

close to the FAV. Species mean ACRs clearly increased as 

Species n Slope R2 P (significance) 95-percent 

confidence limits

Ceriodaphnia dubia1 7 0.441 0.37 0.14 -0.02, 1.10

Daphnia magna, all data 8 .206 .007 .84 -2.3, 2.7

Hyalella azteca, all data2 7 .657 .66 .03 0.11, 1.19

Fathead minnow, all data3 11 .627 .45 .02 0.15, 1.15

1

Not included in covariance analysis, because all but one data point were derived over a narrow range of hardness values. Jop and others 

(1995) data point was excluded because anomalous (≈10 times higher than other values obtained at higher hardnesses).

2

Calculated including Borgmann and others (1991) outlying data point. The Hyalella toxicity-hardness relation used in table 4 to estimate a 

species toxicity–hardness relation excluded this data point, based on the otherwise high degree of agreement of data within this species.

3

Not included in covariance analysis, because several data points were derived over a narrow range of hardness values. Also suspect that 

pooling disparate methods contributes to variability of data.

Table 5. Other data on chronic toxicity values versus hardness not included in individual and pooled regressions.

[n: sample size. Slope: slope of the relations between the natural logarithm (ln) of hardness values and ln toxicity values. R2: coefficient 

of determination or fraction of the variance explained by the regression (Helsel and Hirsch, 2002) P: probability that the apparent relations 

between toxicity and hardness actually resulted from chance alone and that the actual slope of the linear regression is zero. Underlining 

emphasizes P-values that are significant at 0.05 or less]

22 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common 

name Hardness Acute value

Chronic 

value ACR SMACR SMAV SMAV/ FAV Reference

Lepomis macrochirus Bluegill 207 21,100 49.8 423.69 423.69 6,053.63 4,032.06 U.S. Environmental 

Protection Agency, 2001

Jordanella floridae American 

flagfish

44 2,500 5.76 434.03 434.03 2,943.02 1,960.22 U.S. Environmental 

Protection Agency, 2001

Ambystoma gracile Northwestern 

salamander

45 468.4 23.89 19.60 19.60 468.40 311.98 Nebeker and others, 1994, 

1995

Ephemerella2 Mayfly 44 238 1.5 158.67 158.67 238 158.52 Spehar and others, 1978a

Physa integra2 Snail 44 238 5 47.60 47.60 238 158.52 Spehar and others, 1978a

Aplexa hypnorum1 Snail 45.3 93 3.26 28.53 101.01 67.28 Holcombe and others, 1984

Aplexa hypnorum Snail 45.3 93 1.94 47.87 36.95 101.01 67.28 Holcombe and others, 1984

Daphnia pulex Water flea 58 92 7.07 13.01 13.01 73.65 49.06 Ingersoll and Winner, 1982

Ceriodaphnia dubia Zooplankton 44 27.3 2.2 12.41 28.35 18.88 Spehar and Fiandt, 1986

Ceriodaphnia dubia Zooplankton 17 63.1 2 31.5 19.77 28.35 18.88 Suedel and others, 1997

Pimephales promelas Fathead 

minnow

44 13.2 10 1.32 16.49 10.98 Spehar and Fiandt, 1986

Pimephales promelas Fathead 

minnow

201 5,995 63.80 93.97 16.49 10.98 Pickering and Gast, 1972

Pimephales promelas4 Fathead 

minnow

201 5,995 39.23 152.82 16.49 10.98 Pickering and Gast, 1972

Pimephales promelas Fathead 

minnow

17 4.8 1.41 3.39 2.12 16.49 10.98 Suedel and others, 1997

Daphnia magna Water flea 53 9.9 .1523 65.00 14.08 9.38 Chapman and others, 1980

Daphnia magna Water flea 103 33 .2117 155.91 14.08 9.38 Chapman and others, 1980

Daphnia magna Water flea 209 49 .4371 112.09 104.34 14.08 9.38 Chapman and others, 1980

Hyalella azteca Amphipod 17 2.8 .16 17.50 17.50 5.01 3.34 Suedel and others, 1997

Etheostoma fonticola Fountain darter 266 8.68 7.9599 1.09 3.32 2.21 Castillo and Longley, 2001

Etheostoma fonticola Fountain darter 257 14.23 7.953 1.79 3.32 2.21 Castillo and Longley, 2001

Etheostoma fonticola Fountain darter 268 15.7 1.9799 7.93 3.32 2.21 Castillo and Longley, 2001

Etheostoma fonticola Fountain darter 280 13.27 4.6615 2.85 3.32 2.21 Castillo and Longley, 2001

Etheostoma fonticola Fountain darter 278 13.27 6.8367 1.94 2.43 3.32 2.21 Castillo and Longley, 2001

Oncorhynchus 

tshawytscha

Chinook 

salmon

25 1.41 1.12 1.26 1.26 2.67 1.78 Chapman, 1975, 1982

Cottus bairdi Mottled sculpin 100 2.90 .88 3.31 2.56 1.70 Besser and others, 2007

Cottus bairdi Mottled sculpin 100 3.6 1.91 1.89 2.50 2.56 1.70 Besser and others, 2007 

Salmo trutta Brown trout 37.6 2.37 .70 3.39 2.61 1.74 Davies and Brinkman, 1994c

Salmo trutta5 Brown trout 29 1.23 1.02 1.21 2.61 1.74 Brinkman and Hansen, 2007

Salmo trutta5 Brown trout 68 3.9 1.83 2.13 2.61 1.74 Brinkman and Hansen, 2007

Salmo trutta5 Brown trout 151 10.1 6.54 1.55 2.61 1.74 Brinkman and Hansen, 2007

Salmo trutta6 Brown trout 29 1.23 3.52 .35 2.61 1.74 Brinkman and Hansen, 2007

Salmo trutta6 Brown trout 68 3.9 6.36 .61 2.61 1.74 Brinkman and Hansen, 2007

Salmo trutta6 Brown trout 151 10.1 13.56 .75 1.12 2.61 1.74 Brinkman and Hansen, 2007

Table 6. Cadmium acute-chronic ratios (ACRs), ranked in order of closeness of the species mean acute values (SMAV) to the final 

acute value (FAV).

[Underlined values were used to calculate the final ACR. SMACR, species mean acute-chronic ratio]

Results 23
SMAVs increased (R2

 = 0.79, P <0.01). However, selecting a 

cutoff for which SMAVs are suitably close to the FAV was less 

clear. SMAVs that were within about one order of magnitude 

of the FAV were considered “close” to the FAV. This resulted 

in a final ACR value that was close to the value obtained 

from simply using an acute-criterion to chronic-criterion 

ratio (about 3.5). Although the final ACR estimates were not 

used in the criteria derivation, they are presented because 

they might be useful for other purposes, such as site-specific 

modifications of criteria.

Hardness-Based Cadmium Criteria Equations

With the development of hardness-toxicity regressions, 

sufficient information is available to calculate hardnessdependent criterion maximum concentration (CMC or “acute 

criterion”) and the criterion continuous concentration (CCC 

or “chronic criterion”) equations. The criteria derivation 

process involved five major steps: (1) compile and review 

data; (2) develop hardness-toxicity regressions; (3) determine 

genus mean acute and chronic values (GMAVs and GMCVs) 

and rank in order by sensitivity; (4) calculate final acute and 

chronic values (FAVs and FCVs); and (5) derive the hardnessdependent CMC and CCC equations using the hardnesstoxicity slopes.

The results of step 1 were shown in tables 15 and 16, step 

2 in tables 3 and 4, and ranked genus mean acute and chronic 

values (step 3) are shown in table 7. Of these ranked values, 

only the lowest four with cumulative probabilities closest to 

0.05 were used with equation 1 to calculate the FAV (Stephan 

and others, 1985; Erickson and Stephan, 1988). 

Solving equation 1, the FAV was calculated as 

2.451 μg/L cadmium (table 8). This calculated FAV is higher 

than the species mean acute values (SMAVs) for cutthroat 

trout, rainbow trout, and bull trout (table 7). Since these 

species are the Idaho state fish, a popular game fish, and 

a threatened species, respectively, these three species are 

considered “important” species as described in the “Methods” 

section. The calculated FAV is therefore lowered to the lowest 

SMAV of these important species, 1.5 μg/L (Stephan and 

others, 1985, p. 26). 

Species Common 

name Hardness Acute value

Chronic 

value ACR SMACR SMAV SMAV/ FAV Reference

Salvelinus confluentus Bull trout 30 0.95 0.55 1.73 1.73 2.13 1.42 Hansen and others, 2002a; 

2002b

Oncorhynchus mykiss Rainbow trout 21 .84 .88 .95 2.04 1.36 Mebane and others, 2008

Oncorhynchus mykiss Rainbow trout 29 .89 1.58 .56 2.04 1.36 Mebane and others, 2008

Oncorhynchus mykiss3 Rainbow trout 50 2.85 1.47 1.94 2.04 1.36 Davies and others, 1993

Oncorhynchus mykiss3 Rainbow trout 29 2.09 1.39 1.51 1.12 2.04 1.36 Davies and Brinkman, 1994b

Aggregate (FAV to FCV 

ratio)7

N/A 50 1.50 .39 3.85

Final ACR8 3.42

1

 Acute tests were conducted with adult snails. With some species, adults are more acutely resistant to cadmium than juveniles. Thus, ACR estimates using 

adults might be high biased. 

2

 Qualitative estimate, acute value is “greater than,” chronic value is estimated from authors graphs.

3

 Using average of two acute tests conducted concurrent to chronic test using same dilution water.

4

 Acute tests conducted with adult fish which were much more resistant to cadmium than newly hatched fry. ACR estimates considered invalid. 

5

 ACR calculated using juvenile growth and survival test as denominator.

6

 ACR calculated using early life stage test as denominator.

7

 Calculated as an estimate of overall ACR values, to provide a separate comparison of species-by-species ACRs. In cases where the species mean ACRs 

seem to increase as the SMAVs increase, Stephan and others (1985) recommends that the final ACR be a calculated species which SMAVs are close to the FAV.

8

 Geometric mean of the species mean acute-chronic ratios (SMACRs) from species for which their SMAVs were within about an order of magnitude of the 

FAV (actually 11 times difference, using underlined SMACR values).

Table 6. Cadmium acute-chronic ratios (ACRs), ranked in order of closeness to species mean acute values (SMAV) and final acute 

value (FAV).—Continued

24 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 7. Ranked genus mean acute and chronic values. 

[Ranked in order of least resistant to most resistant. Abbreviations: GMAV, genus mean acute value; GMCV; genus mean chronic value; SMAV, species mean acute value; SMCV, species mean chronic value. Symbols: mg/L, microgram per liter; –, missing or inapplicable values]

Acute ranking

GMAV 

rank

GMAV 

(μg/L)

SMAV 

(μg/L) Family Species Common name

1 2.02 1.50 Salmonidae Oncorhynchus clarki Cutthroat trout1 2.02 2.04 Salmonidae Oncorhynchus mykiss Rainbow trout1 2.02 2.67 Salmonidae Oncorhynchus tshawytscha Chinook salmon2 2.13 2.13 Salmonidae Salvelinus confluentus Bull trout3 2.61 2.61 Salmonidae Salmo trutta Brown trout4 2.61 2.56 Cottidae Cottus bairdi Mottled sculpin4 2.61 2.67 Cottidae Cottus confusus Shorthead sculpin5 3.32 3.32 Percidae Etheostoma fonticola Fountain darter6 5.39 5.39 Hyalellidae Hyalella azteca Amphipod7 8.29 8.29 Salmonidae Prosopium williamsoni Mountain whitefish8 16.5 16.5 Cyprinidae Pimephales promelas Fathead minnow9 22.7 22.7 Pontoporeiidae Diporeia sp. Amphipod

10 26.2 26.2 Daphniidae Simocephalus serrulatus Cladoceran11 28.6 12.2 Daphniidae Daphnia ambigua Cladoceran11 28.6 22.3 Daphniidae Daphnia magna Cladoceran11 28.6 87.1 Daphniidae Daphnia pulex Cladoceran12 32.4 45.8 Unionidae Lampsilis straminea 

claibornensis

Southern fatmucket 

(mussel)

12 32.4 23.0 Unionidae Lampsilis teres Yellow sandshell 

(mussel)

13 32.9 32.9 Daphniidae Ceriodaphnia dubia Cladoceran14 36.9 36.9 Unionidae Actinonaias pectorosa Pheasant shell 

(mussel)

15 38.8 38.8 Unionidae Villosa vibex Southern rainbow 

(mussel)

16 45.7 45.7 Unionidae Utterbackia imbecilis Paper pondshell 

(mussel)

17 47.8 66.8 Cambaridae Orconectes juvenilis Crayfish17 47.8 34.2 Cambarinae Orconectes placidus Placid crayfish

Chronic ranking

GMCV 

rank

GMCV 

(μg/L)

SMCV 

(μg/L) Family Species Common name

1 0.33 0.33 Hyalellidae Hyalella azteca Amphipod

2 .84 .84 Cottidae Cottus bairdi Mottled sculpin

3 1.00 1.00 Gammaridae Gammarus fasciatus Amphipod4 1.25 2.01 Salmonidae Salvelinus fontinalis Brook trout

4 1.25 .77 Salmonidae Salvelinus confluentus Bull trout5 1.41 .54 Daphniidae Daphnia magna Cladoceran5 1.41 3.70 Daphniidae Daphnia pulex Cladoceran

6 1.62 .87 Salmonidae Salmo salar Atlantic salmon6 1.62 3.03 Salmonidae Salmo trutta Brown trout

7 1.81 1.81 Percidae Etheostoma fonticola Fountain darter8 1.87 2.28 Salmonidae Oncorhynchus 

kisutch

Coho salmon

8 1.87 1.68 Salmonidae Oncorhynchus 

mykiss

Rainbow trout

8 1.87 1.72 Salmonidae Oncorhynchus 

tshawytscha

Chinook salmon

9 2.21 2.21 Bufonidae Bufo americanus American toad

10 2.68 2.68 Physidae Aplexa hypnorum Snail 

11 3.32 3.32 Chironomidae Chironomus tentans Midge12 3.36 2.04 Daphniidae Ceriodaphnia dubia Cladoceran

12 3.36 5.63 Daphniidae Ceriodaphnia 

reticulata

Cladoceran

13 5.32 5.32 Cyprinodontidae Jordanella floridae American flagfish

14 6.33 6.33 Cyprinidae Pimephales promelas Fathead minnow

15 7.69 7.69 Catostomidae Catostomus 

commersoni

White sucker

16 7.97 7.97 Esocidae Esox lucius Northern pike

Results 25
Table 7. Ranked genus mean acute and chronic values. —Continued 

Acute ranking

GMAV 

rank

GMAV 

(μg/L)

SMAV 

(μg/L) Family Species Common name

18 59.2 59.2 Asellidae Lirceus alabamae Isopod19 76.7 76.7 Gammaridae Gammarus pseudolimnaeus Amphipod20 81.7 96.3 Hydridae Hydra oligactis Hydra20 81.7 63.2 Hydridae Hydra viridissima Hydra20 81.7 89.6 Hydridae Hydra vulgaris Hydra21 101 101 Physidae Aplexa hypnorum Physid snail22 103 103 Heptageniidae Rhithrogena sp. Mayfly23 129 129 Physidae Physa gyrina Physid snail24 151 151 Planorbidae Gyraulus sp. Gyro snail25 179 179 Lumbriculidae Lumbriculus variegatus Oligochaete worm26 193 193 Lymnaeidae Lymnaea stagnalis Pulmonate pond snail, 

swamp lymnaea

27 226 226 Hydrobiidae Potamopyrgus antipodarum New Zealand mud 

snail

28 226 226 Glossiphoniidae Glossiphonia complanata Leech29 306 306 Baetidae Baetis tricaudatus Mayfly30 354 354 Salmonidae Coregonus clupeaformis Lake whitefish31 512 512 Ambystomatidae Ambystoma gracile Northwestern 

salamander

32 616 616 Asellidae Asellus sp. Isopod33 716 716 Arctopsychidae Arctopsyche sp. Caddisfly34 722 722 Cambaridae Procambarus clarkii Red swamp crayfish 35 832 832 Cyprinidae Carassius auratus Goldfish36 1,026 1,026 Tubificidae Limnodrilus hoffmeisteri Tubificid worm37 1,332 1,332 Tubificidae Tubifex tubifex Tubificid worm38 1,346 1,346 Aeolosomatidae Aelosoma headleyi Polychaete worm39 1,570 1,570 Tubificidae Branchiura sowerbyi Tubificid worm40 1,950 1,950 Cyprinidae Ptychocheilus oregonensis Northern pikeminnow41 2,093 2,093 Tubificidae Quistadrilus multisetosus Tubificid worm42 2,485 2,485 Tubificidae Varichaetadrilus pacificus Tubificid worm43 2,289 2,289 Tubificidae Spirosperma sp. (“ferox”) Tubificid worm

Chronic ranking

GMCV 

rank

GMCV 

(μg/L)

SMCV 

(μg/L) Family Species Common name

17 8.00 8.00 Centrarchidae Micropterus 

dolomieui

Smallmouth bass

18 20.5 20.5 Centrarchidae Lepomis macrochirus Bluegill19 23.3 23.3 Aeolosomatidae Aelosoma headleyi Polychaete worm20 25.5 25.5 Ambystoma Ambystoma gracile Northwestern 

salamander

21 26.7 26.7 Cichlidae Oreochromis aurea Blue tilapia

26 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Acute ranking

GMAV 

rank

GMAV 

(μg/L)

SMAV 

(μg/L) Family Species Common name

43 2,595 2,943 Tubificidae Spirosperma nikolskyi Tubificid worm44 2,610 2,610 Catostomidae Catostomous commersoni White sucker45 2,617 2,617 Poeciliidae Poecilia reticulata Guppy46 2,782 2,782 Cyprinodontidae Jordanella floridae American flagfish47 3,597 3,597 Lumbriculidae Stylodrilus heringlianus Oligochaete worm48 3,913 3,913 Percidae Perca flavescens Yellow perch49 4,120 4,120 Tubificidae Rhyacodrilus montana Tubificid worm50 4,226 4,226 Cyprinidae Cyprinella lutrensis Red shiner51 4,986 4,986 Ictaluridae Ictalurus punctatus Channel catfish52 4,995 4,995 Poeciliidae Gambusia affinis Mosquitofish53 5,791 5,540 Centrarchidae Lepomis cyanellus Green sunfish53 5,791 6,054 Centrarchidae Lepomis macrochirus Bluegill54 7,866 7,866 Perlodidae Perlodidae Stonefly55 11,742 6,988 Chironomidae Chironomus riparius Midge55 11,742 19,730 Chironomidae Chironomus tentans Midge56 12,328 12,328 Gasterosteidae Gasterosteus aculeatus Threespine 

stickleback

57 15,540 15,540 Dendrocoelidae Dendrocoelum lacteum Planarian

Acute values:Number of species: 69Number of genera: 57Number of families: 33

Table 7. Ranked genus mean acute and chronic values. —Continued 

Chronic ranking

GMCV 

rank

GMCV 

(μg/L)

SMCV 

(μg/L) Family Species Common name

Chronic data:

Number of species: 28Number of genera: 21Number of families: 17

Results 27
Table 8. Acute and chronic criteria calculations.

[See equation 1. Abbreviations: GMAV, genus mean acute value; SQR, (square root); n, number; FAV, final acute value; FCV, final chronic value. mg/L, 

microgram per liter]

Solution to the FAV equation using a column and row spreadsheet matrix format 

(Syntax shown is for the Excel® spreadsheet program (Microsoft Corporation, Redmond, Wash.)

A B C D E F G

1 Genus Rank GMAV Cumulative

 Probability

 (P=R/(n+1)

ln GMAV (ln GMAV)2 SQRT P

2 Cottus 4 2.610 =B2/(B$6+1) =LN(C2) =POWER(E2,2) =SQRT(F2)

3 Salmo 3 2.610 =B3/(B$6+1) =LN(C3) =POWER(E3,2) =SQRT(F3)

4 Salvelinus 2 2.126 =B4/(B$6+1) =LN(C4) =POWER(E4,2) =SQRT(F4)

5 Oncorhynchus 1 2.017 =B5/(B$6+1) =LN(C5) =POWER(E5,2) =SQRT(F5)

6 n= 57

7 S= =SQRT(C7) =(F7-E7)/ 

 (D7-G7)

=SUM(D2:D5) =POWER(SUM

 (E2:E5),2)/4

=SUM(F2:F5) =(POWER(SUM

(G2:G5),2))/4

8 L= =(SUM(E2:E5)

 -C8)/4

=(B7*SUM 

(G2:G5))

9 A= =(B7*SQRT

 (0.05))+B8

10 FAV= =EXP(B9)

11 CMC= B10/2

Final acute value (FAV) equation results

Rank Genera GMAV 

(μg/L)

Cumulative

probability 

(P=R/(n+1)

ln GMAV (ln GMAV)2 SQRT P

4 Cottus (sculpin) 2.610 0.07 0.9595 0.9206 0.2626

3 Salmo (Atlantic trout and salmon) 2.610 .05 .9593 .9203 .2357

2 Salvelinus (char) 2.126 .03 .7542 .5688 .1925

1 Oncorhynchus (Pacific trout and salmon) 2.017 .02 .7017 .4924 .1313

Sum: 0.17 3.3908 2.934 0.8362

Number of GMAVs = 57; S = 2.390; L = 0.3619; A = 0.8962.

FAV (calculated) = 2.450 μg/L.

Calculated FAV lowered to protect cutthroat trout, bull trout and rainbow trout = 1.50 μg/L.

(Lowered to the lowest SMAV for these three “important” species which have SMAVs lower than the FAV (Stephan and others, 1985, p. 26).

CMC = FAV/2 = 0.75 μg/L

Final chronic value (FCV)

Rank Genera GMCV

 (μg/L)

Cumulative 

probability 

(P=R/(n+1)

ln GMCV (ln GMCV)2 SQRT P

4 Salvelinus (char) 1.281 0.18 0.2476 0.06120 0.4264

3 Gammarus (amphipod) 1.004 .14 3.501E-3 1.226E-05 .3692

2 Cottus (sculpin) .838 .09 -0.1763 .03109 .3015

1 Hyalella (amphipod) .326 .05 -1.124 1.263 .2132

Sum: 0.4545 -1.0758 1.343 1.310

Number of GMCVs = 21; S = 6.458; L = -2.384; A = -0.940.

FCV = 0.39 μg/L.

28 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
The FCV was calculated similarly using equation 2 as 

shown in table 8. The species mean chronic value (SMCV) for 

one species, Hyalella azteca, was lower than the FCV (tables 7 

and 8). Unlike the FAV derivation, the FCV was not lowered to 

explicitly protect Hyalella azteca because Hyalella azteca was 

not obviously an important or critical species as defined in the 

“Methods” section. Whether the toxicity of cadmium to this 

species at FCV concentrations would likely result in indirect 

effects to other species or alter benthic assemblages such that 

biological integrity could be compromised are considered in 

more detail in the section “Risks of the 5th Percentile SpeciesSensitivity Distribution (SSD)-Based Chronic Criterion to a 

More Sensitive Species.”

The CMC is the FAV/2 (table 8). The hardness-based 

CMC equation has two components, the slope and the criterion 

maximum intercept (CMI). The slope is the same as that 

derived in step 2 (table 3). The intercept was calculated as a 

function of the natural log (ln) of the hardness-specified CMC 

and the slope. The intercept equation is:

ln(CMI) = ln(hardness-specified CMC) 

– (slope  ́ ln(specified hardness))

Once the intercept is determined, the CMC equation is:

CMC at hardness X = e((slope  ́ ln(X)) + ln(CMI))

Using the CMC and hardness slope from tables 3 and 8, 

the hardness-based CMC equation is:

CMC (μg/L dissolved cadmium) = e(0.8403  ́ ln(hardness)-3.572)

where the (ln hardness) is the natural logarithm of the water 

hardness. 

The CCC hardness-dependent equation was similarly 

derived. Two differences between the CMC and CCC 

derivation were that first, unlike the FAV, the FCV was not 

divided by two because the chronic values used to develop 

the FCV reflect less severe effects than 50 percent mortality. 

Second, because of the belief that dissolved metals (the 

fraction of metals that pass a 0.45-μm filter) better predict 

toxicity than total metals concentrations from unfiltered 

samples, and because most chronic values used to derive the 

CCC were derived with unfiltered samples, an estimated totalto-dissolved cadmium conversion factor is used with the CCC.

The chronic conversion factor (CF) used was 

CF = 1.101672 – ((ln hardness)  ́ 0.041838)), 

which at a hardness of 50 mg/L equals 0.94 (Stephan, 1995). 

No conversion for the CMC was needed since the FAV was 

based tests that reported dissolved values. The CCC hardnessdependent equation is 

CCC (μg/L dissolved cadmium) = (e(0.6247 ́ln(hardness)-3.384)) 

 ́ (1.101672 – ((ln hardness)  ́ 0.041838)). 

These acute and chronic cadmium criteria values are 

compared to previous cadmium criteria derivations in table 9.

, (1)

where

n is number of genera with Genus Mean Acute Values 

(GMAVs),

R is relative sensitivity rank of each GMAV, 

P is cumulative probability for each GMAV.

, (2)

where

n is number of genera with Genus Mean Chronic Values 

(GMCVs),

R is relative sensitivity rank or each GMCV, and

P is cumulative probability for each GMCV.

Results 29

P R n

S

GMCV GMCV

(P) (( ( P

= +

= −

−

∑ ∑

∑ ∑

/( )

(( ( / )

)) / )

1

4

4

2

2

ln ) ( ( ln ))2

L GMCV S P

A S L

FCV eA

= ( − ( ) ( ) )

= +

=

∑(ln ) ∑ /

( . )

4

0 05

P R n

S

GMAV GMAV

(P) (( ( P

= +

= −

−

∑ ∑

∑ ∑

/( )

(( ( / )

)) / )

1

4

4

2

2

ln ) ( ( ln ))2

L GMAV S P

A S L

FAV e

A

= ( − ( ) ( ) )

= +

=

∑(ln ) ∑ /

( . )

4

0 05
Aquatic Toxicology of Cadmium—Synthesis

Mechanisms and Factors Affecting Aquatic 

Toxicology of Cadmium

Waterborne cadmium can cause severe, acute 

toxicological and physiological effects to aquatic organisms. 

At acute waterborne concentrations, cadmium severely 

disrupts calcium homeostasis, which ultimately leads to 

death. At the gill surface, cadmium competes with calcium 

for high affinity calcium-binding sites and once it enters 

the chloride cell irreversibly blocks calcium uptake. The 

cumulative effect of these two processes causes acute 

hypocalcaemia in freshwater fish (Wood and others, 1997). 

However, these effects can be altered by the calcium content 

of the water. Among various water-quality parameters that 

often influence the uptake and toxicity of metals, calcium 

seems to be of primary importance, with pH of secondary 

importance. Other water-quality parameters that influence 

the toxicity of other metals (such as copper) probably are 

less important in modifying cadmium toxicity (dissolved 

organic matter (DOM), magnesium, sodium, and alkalinity). 

This protective action of calcium has been attributed to 

changes in gill permeability and (or) competition between 

cadmium and calcium for gill-binding sites. Toxicity is 

reduced because calcium out-competes cadmium for binding 

sites on gills. Protection from cadmium uptake and acute 

toxicity in freshwater fish is related to calcium concentration 

in water rather than magnesium or alkalinity concentrations 

indicating that calcium is the primary cation responsible 

for the protective action of hard water (Carroll and others, 

1979; Davies and others, 1993). In many tests, water 

hardness (defined as the sum of calcium and magnesium 

concentrations, expressed as CaCO3 equivalents) is often 

reported in lieu of calcium concentrations. Because calcium 

is the principle cation of hardness in most natural waters, 

hardness is often a reliable surrogate measure of calcium 

(Hem, 1992). Exceptions to this generalization include some 

synthetic laboratory test waters or atypical natural waters, 

where the presumption that hardness is a good surrogate for 

the protective effects of calcium would not hold (Welsh and 

others, 2000).

Moreover, the bioavailability and toxicity of cadmium 

are less affected by DOM than copper because the binding 

affinity of cadmium to the gill is stronger than the binding 

affinity of cadmium to DOM (Winner, 1984; Playle and 

others, 1993; Playle, 2004). DOM from some natural 

Source1 Number of test 

values used

Number of genera 

used to calculate SSD

 5th percentile value (n)

Criterion value 

calculated from 

5th percentile of SSD

Final criterion value 

(μg/L total Cd)

Acute criteria derivation

This report1 277 57 1.2 0.75

U.S. Environmental Protection 

Agency, 20012 226 55 1.4 1.0

U.S. Environmental Protection 

Agency, 19843 133 44 4.4 1.8

Chronic criteria derivation

This report 93 21 0.39 0.39

U.S. Environmental Protection 

Agency, 2001 34 16 .17 .17

U.S. Environmental Protection 

Agency, 19844 25 44 .66 .66

1

The final acute criterion value was lowered to protect cutthroat trout, rainbow trout and bull trout, important species with species mean acute values lower 

than the 5th percentile of the SSD of GMAVs.

2

The final acute criterion value was lowered to protect rainbow trout and bull trout.

3

The final acute criterion value was lowered to protect rainbow trout.

4

The actual number of genera in dataset was 13, not 44. However, because the 13 mean genus chronic values contained values for 5 of the 6 genera most 

acutely sensitive to cadmium, it was judged more appropriate to calculate the FCV using 44 genera instead of 13 genera. The effect of this judgment was to 

raise the FCV from 0.04 μg/L to 0.66 μg/L cadmium (U.S. Environmental Protection Agency, 1984, p. 10, 46).

Table 9. Comparison of acute and chronic cadmium criteria values derived in this report to previously reported cadmium criteria.

[Criteria values all calculated for a hardness of 50 milligrams per liter; cadmium values are for total (unfiltered) concentrations. Abbreviations: SSD, species 

sensitivity distribution; Cd, cadmium; GMAV, genus mean acute value; FCV, final chronic value. Symbols: mg/L, microgram per liter]

30 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
sources made cadmium more toxic to fish, whereas copper 

and lead were less toxic in the presence of DOM (Schwartz 

and others, 2004). Calcium is therefore a more important 

modifier of cadmium toxicity in fish than water DOM, which 

plays an important role in modifying the toxicity of some 

metals, such as copper and lead. Water pH also modifies 

cadmium toxicity, with increasing cadmium toxicity with 

increasing pH (Campbell and Stokes, 1985; Cusimano and 

others, 1986; Schubauer-Berigan and others, 1993; Hansen 

and others, 2002a). Because in natural and synthetic waters 

where pH is not artificially manipulated water hardness is 

usually correlated with pH, water hardness probably is a 

good surrogate for the specific factors affecting cadmium 

bioavailability and toxicity. For example, using EPA’s recipes 

for preparing synthetic freshwaters, very soft water with a 

hardness of about 10–13 mg/L will have pH values ranging 

from about 6.4 to 6.8, and very hard water with a hardness of 

about 280–320 mg/L will have pH values ranging from about 

8.0 to 8.4 (U.S. Environmental Protection Agency, 2002c).

Toxicity of Cadmium in Chemical Mixtures

Cadmium probably seldom occurs as a single 

contaminant of concern in releases to ambient waters, but 

commonly occurs with zinc and copper. Evaluating interactive 

effects of even binary mixtures is a complex and difficult topic 

in ecological risk assessment. The joint toxicity of chemicals 

is often described as synergism, antagonism, or additivity. 

The toxicity of a particular chemical when combined with 

another can remain unchanged (additive responses), increase 

(synergism, or more than additive responses), or decrease 

(antagonism, or less than additive responses) (Sorensen, 1991; 

Hertzberg and MacDonell, 2002; Norwood and others, 2003). 

Although a detailed treatment of the joint toxicity of cadmium 

with other chemicals is beyond the scope of this report, the 

following results are relevant.

Although cadmium, copper, and zinc have some 

toxicological similarities, the toxicity of cadmium to fish in 

the presence of zinc or copper seems to result in only slightly 

increased toxicity over that of cadmium alone. In acute tests of 

cadmium plus zinc mixtures with bull trout and rainbow trout, 

cadmium and zinc LC50s were similar when tested singly or 

as a mixture (Hansen and others, 2002a). Similarly in chronic 

toxicity tests of cadmium plus zinc with the American flagfish, 

no- and lowest-observed-effect concentrations of cadmium and 

zinc were similar when tested singly or as mixtures (Spehar 

and others, 1978b). The acute LC50s obtained with Chinook 

salmon from cadmium singly or with two and three metal 

mixtures (cadmium + copper, and cadmium + copper + zinc) 

were slightly lower with mixtures than with cadmium singly 

in some combinations of ratios. However, the differences were 

not statistically significant and the three metals could have 

been exhibiting independent action (Finlayson and Verrue, 

1982). 

Contradictory patterns were observed with invertebrates 

in toxicity testing and field experiments. In 96-hour tests with 

the paper pondshell mussel Utterbackia imbecilis, Keller and 

Zam (1991) found that cadmium was slightly less toxic in 

combination with copper, indicating that copper and cadmium 

were slightly antagonistic to each other. Shaw and others 

(2006) tested the joint toxicity of cadmium and zinc mixtures 

to four species of zooplankton. Mixture toxicity was governed 

primarily by zinc. The presence of zinc at concentrations that 

caused low toxicity (LC15) tended to reduce cadmium toxicity. 

However, when the exposure was reversed to test a range of 

zinc concentrations in the presence of cadmium at its LC15, no 

interaction was indicated between the metals (no increase or 

decrease in zinc toxicity). Attar and Maly (1982) also noticed 

that cadmium and zinc mixtures were less toxic than expected. 

In contrast, in stream microcosm and field experiments, 

Clements (2004) observed that macroinvertebrate responses to 

a mixture of copper, cadmium, and zinc were greater than to 

cadmium plus zinc mixtures or zinc alone.

In urban or industrial settings, metals and polynuclear 

aromatic hydrocarbons often co-occur in runoff. Gust (2006) 

investigated the joint toxicity of cadmium and phenanthrene 

to the amphipod Hyalella azteca in various waterborne and 

sediment exposures. The mixtures could cause synergistic or 

independent toxicity in Hyalella depending on the endpoint 

investigated and the experimental protocol employed. Adding 

sublethal phenanthrene concentrations to sediment decreased 

cadmium LC50s, that is, increased the risk of cadmium lethality 

to Hyalella. Yet adding sublethal phenanthrene concentrations 

to lower sublethal cadmium concentrations had no effect on 

sublethal cadmium effects thresholds to Hyalella. In acute 

waterborne exposures, the presence of sublethal phenanthrene 

concentrations decreased the toxicity of cadmium. 

Generally, these examples indicated that the toxicity of 

cadmium is little changed in the presence of other metals. In a 

more comprehensive review of metals mixture toxicity studies, 

Norwood and others (2003) reported a tendency toward “less 

than additive” results in metals mixture testing. However, in a 

minority of published studies of the joint toxicity of cadmium 

in binary mixtures (15 of 53 or 28 percent), investigators 

reported “greater than additive effects” (synergism). This 

suggests that in situ assessments of toxicity may be prudent 

if joint toxicity of metals mixtures are of concern, such as 

benthic community monitoring or laboratory testing of fieldcollected water. 

Relative Sensitivity of Chronic Test Endpoints

Some generalizations about the relative sensitivity of 

different chronic endpoints in different species are supported 

by the data reviewed. In chronic tests with salmonids, the 

principle adverse effect was most often acute mortality. In 

several chronic tests with cadmium and salmonids, the onset 

of increased mortalities began shortly after the newly hatched 

fish reached swim-up stage, with mortalities then ceasing 

or greatly diminishing by about 4–5 days after their onset. 

Results 31
The fish that survived this acute phase usually survived in 

the test concentrations for several more weeks until the tests 

ended with subtle if any apparent adverse effects (Hollis and 

others, 1999, 2000b; Hansen and others, 2002b; Mebane and 

others, 2008). Similarly, in 3 pairs of 4- and 30-day exposures 

of brown trout to cadmium, Brinkman and Hansen (2007) 

obtained similar LC50s from the 4- or 30-day tests (table 17). 

In contrast to the similarity in acute and chronic sensitivities 

of salmonids, 28-day exposures of newly hatched mottled 

sculpin produced more sensitive results than did 4-day 

exposures (tables 15 and 16). Growth of juveniles seldom 

was the most sensitive endpoint with cadmium exposures in 

tests reviewed with fish, and in few tests with invertebrates. 

Reproduction often was the most sensitive endpoint in tests 

with invertebrates.

Bioaccumulation and Effects of Dietary Exposures to 

Cadmium

Most analyses presented in this report focus on relating 

observed effects in organisms to concentrations of cadmium 

in water to estimate thresholds of acceptable risk for those 

organisms. However, this approach is less direct than relating 

observed effects to concentrations of cadmium in the target 

tissues of the organism because a contaminant cannot be 

toxic to an animal if it is not accumulated in its tissues. For 

example, an acute LC50 or chronic LOEC water concentration 

is simply a surrogate for the amount of the toxicant in the 

organism at the site(s) of toxic action producing the observed 

mortality. Additionally, organisms may be exposed to trace 

metals such as cadmium through food or sediments even 

when cadmium concentrations in the water column are low. 

Thus the consequences of elevated tissue residues or effects 

of dietary exposures may be important when estimating 

protective thresholds for cadmium and other pollutants 

(McCarty and Mackay, 1993; Meyer and others, 2005). 

At least on a whole-body bioaccumulation basis, for some 

organisms cadmium uptake from prey is as important or more 

important than uptake from water. Stephenson and Turner 

(1993) examined the relation between dietary uptake of 

cadmium from periphyton and bioaccumulation in Hyalella 

azteca in a cadmium contaminated lake and a control lake. 

Dietary sources accounted for about 58 percent of cadmium 

accumulated. To determine the relative importance of food and 

water as cadmium sources for benthic insects, Roy and Hare 

(1999) measured cadmium accumulation by the predatory 

alderfly Sialis velata (Megaloptera) from either water alone or 

from chironomid prey. There, cadmium uptake from prey was 

far more important than that from water (Roy and Hare, 1999).

Further, if a maximum acceptable dietary intake or tissue 

concentration for fish or other aquatic-dependent life such as 

amphibians could be demonstrated based on observations of 

survival, growth, or reproduction in a chronic feeding study or 

a long-term field study, a “final residue value” criterion based 

upon tissue residues and bioaccumulation or bioconcentration 

factors could be defined (Stephan and others, 1985). More 

generally, the term “tissue reference value” is sometimes used 

for tissue residue concentrations above which adverse effects 

in aquatic biota may occur. The following data summary 

highlights selected studies that were relevant to the discussion 

of toxicity associated with cadmium tissue residues and were 

relevant to coldwater food chains (table 10). See Jarvinen 

and Ankley (1999) and Bridges and Lutz (2005) for more 

encyclopedic summaries.

Fish

At least for fish, waterborne cadmium accumulation 

on gill surfaces appears to be a primary site of toxic action. 

Effects associated with gill residues are emphasized here, 

because gill tissue was suggested as the optimum body tissue 

to predict cadmium toxicity and tissue-residue relation for fish 

(Jarvinen and Ankley, 1999). The highest concentrations of 

cadmium in fish usually occurs in the kidney, however it is not 

clear whether cadmium residues in the kidney are bioactive or 

if the residues are simply sequestered in the kidney. Cadmium 

residues in kidneys appear to be permanent and remain 

elevated long after the exposures to elevated cadmium from 

water or food end.

Several examples of cadmium tissue residues that 

corresponded with the absence or presence of observed 

adverse effects are listed in table 10. There was considerable 

overlap in the concentrations associated with the absence or 

presence of observed adverse effects for a given tissue type. 

Rainbow trout were able to tolerate cadmium in gill tissues 

of as much as 38 mg/kg dry weight (dw) without noticeable 

adverse effects, yet complete mortality to brook trout occurred 

in a study in which gill residues were only about 5 mg/kg dw. 

In contrast, brook trout living in a cadmium polluted river in 

Idaho carried an average gill tissue burden of 127 mg/kg dw. 

This 25 times difference between lethal and non-lethal gill 

residues for the same species suggests that it may be difficult 

to predict cadmium toxicity tissue-residue relations for fish.

Most studies with fish indicated that dietary cadmium 

exposures at environmentally realistic concentrations resulted 

in bioaccumulation but no appreciable adverse effects. Only 

when exorbitant concentrations of dietary cadmium were 

administered were obvious adverse effects observed. Scott and 

others (2003) observed that rainbow trout fed only 3 mg/kg 

through diet showed inhibition of cortisol elevation (a stress 

hormone) following exposure to alarm chemicals. However, 

the significance of this hormonal change is unclear because 

no behavioral changes were associated with the inhibition of 

cortisol elevation due to dietary cadmium exposures, unlike 

waterborne exposures which resulted in similar cortisol 

inhibition and similar whole-body cadmium residues (Scott 

and others, 2003). 

32 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 10. Examples of cadmium tissue residues co-occurring with the absence or presence of adverse effects.

[Abbreviations: NOEC, no observed effect concentration; LOEC, lowest observed effect concentration; Cd, cadmium; TRV, tissue reference value. Symbols:

mg/kg, milligram per kilogram; L/kg, liter per kilogram; mg/L, milligram per liter; μg/g, microgram per gram; μg/L, microgram per liter]

Organism

Residue 

concentration 

(mg/kg dry 

weight)

Tissue Effect Source

Aquatic communities 2.9 – 14.6 Whole body TRV a tissue residue in aquatic biota above 

which adverse ecological effects may occur. 

Calculated as the product of the chronic 

water-based criterion over a hardness range 

of 20 – 250 mg/L (0.22 – 1.1 μg/L) and a 

bioconcentration factor of 2623 L/kg

This report

0.4 – 2.0 Whole body TRV, tissue reference value calculated as the 

product of the chronic water-based criterion 

over a hardness range of 20 – 250 mg/L 

(0.22 – 1.1 μg/L) and an accumulation factor 

(352 L/kg) that was adjusted to factor out 

background concentrations 

This report

Tissue residues corresponding with the absence of apparent adverse effects (highest NOECs from each study)

Amphipod (Hyalella 

azteca)

0.1–37 Whole body No apparent effects; range of concentration 

from lakes in central Ontario, Canada with no 

direct pollution sources

Stephenson and Mackie, 

1988

0.32 Diet Low reductions in growth (EC25) in Hyalella

fed a diet of Cd contaminated green algae. 

Corresponding Cd exposure in water 

producing the Cd in algae EC25 was about 

0.57 μg/L cadmium at hardness 125 mg/L. No 

NOEC obtained

Ball and others, 2006

13–62 Whole body No effect on growth, survival, or reproduction Borgmann and others, 

1991

1–3 Whole body No effect; values from control treatments Borgmann and others, 

1991

Atlantic salmon (Salmo 

salar

~0.6 Whole body Minimal reduction in alevin biomass, 

corresponding to a water LOEC of 0.47 μg/L, 

see table 16

Rombough and Garside, 

1982

Brook trout (Salvelinus 

fontinalis)

10 Kidney No effect Benoit and others, 1976

~2.5 Gill No effect Benoit and others, 1976

0.5 Gill No effect; mean value from wild fish collected 

from an unpolluted reference river, North 

Fork Coeur d’Alene River, Idaho

Farag and others, 1998

Crustacea (multiple taxa) <0.3–53 Whole body No apparent effects; range of concentrations 

from a variety of marine sites around the 

world with no direct pollution sources 

Rainbow, 2002

Lake trout (Salvelinus 

namaycush)

20–45 Kidney No apparent effects; concentrations following 

3+ years of whole-lake Cd dosing to near 

chronic criterion levels

Malley, 1996

Longnose dace 

(Rhinichthys cataractae)

3–8 Gill No effect Riddell and others, 2005a

Mallard (Anas 

platyrhynchus)

200 Diet No obvious deleterious effects after 13 weeks Eisler, 1985

Mayfly (Baetis 

tricaudatus)

27 Whole body No apparent effect on population densities Cain and others, 2004

Mayfly (Rhithrogena sp.) 6–34 Whole body No apparent effects; range of concentration 

from least-disturbed Colorado streams with 

apparently healthy fish and invertebrate 

communities)

Besser and others, 2001

Results 33
Organism

Residue 

concentration 

(mg/kg dry 

weight)

Tissue Effect Source

Tissue residues corresponding with the absence of apparent adverse effects (highest NOECs from each study)–Continued

Mayfly (Serratella tibialis) 35 Whole body Apparently near the limit of tolerance since 

this was the highest concentration observed 

in Serratella collected along a gradient of 

contamination. No obvious adverse effects 

were associated with this concentrations

Cain and others, 2004

Rainbow trout 

(Oncorhynchus mykiss)

30 Gill No apparent effect on population density and 

biomass in comparison to reference streams

Farag and others, 2003

7.5 Gill No effect. 99 percent of whole-body Cd was 

accumulated in the gills, liver, and kidney

Thomas and others, 1983

8.5 Gill No effect on survival or growth; residue 

resulting from 28-day exposure to 3 μg/L Cd 

at hardness of 140 mg/L

Franklin and others, 2005

11 Gill No effect on survival or growth; residue 

resulting from 36-day exposure to 2 μg/L Cd 

at hardness of 140 mg/L1

Szebedinszky and others, 

2001

19.5 Gill Saturation concentration following 65 to100-day 

exposures to 3 μg/L Cd at hardness of 140 

mg/L. No effects on growth, survival or 

swimming speed1

McGeer and others, 2000a, 

2000b

94 Kidney Saturation concentration following 65 to100-day 

exposures to 3 μg/L Cd at hardness of 140 

mg/L. No effects on growth, survival or 

swimming speed1

McGeer and others, 2000a, 

2000b

4.75 Whole body Saturation concentration following 65 to100-day 

exposures to 3 μg/L Cd at hardness of 140 

mg/L. No effects on growth, survival or 

swimming speed1

McGeer and others, 2000a, 

2000b

30 Diet (spiked trout 

chow)

No effect on survival or growth; fed Cd spiked 

food for 12-weeks and monitored for an 

additional 6-weeks

Kumada and others, 1973

500 Diet (spiked trout 

chow)

No effect; after 14-days gill residues reached 

5 μg/g dw from dietary exposure (no 

waterborne exposure)

Franklin and others, 2005

786 Diet (spiked trout 

chow)

No effect on survival or growth Szebedinszky and others, 

2001

2.2 Diet (live prey) No effects on growth or survival. Exposed 

via live diet of Lumbriculus, an aquatic 

oligochaete that had been reared in Cd 

contaminated river sediments

Hansen and others, 2004

55 Diet (live prey) No significant effects on growth or survival; 

trout fry were fed live brine shrimp (Artemia

sp.) enriched with Cd

Mount and others, 1994

300 Diet (spiked trout 

chow)

No significant effects on growth, survival, or 

total body calcium accumulation

Baldisserotto and others, 

2005

Snail (Potamopyrgus 

antipodarum)

7.5 Whole body (without 

shell)

No significant decreases in population growth 

rate of clones from different areas of northern 

Europe or New Zealand 

Jensen and others, 2001

Zooplankton (Daphnia 

galeata mendotae)

28.3 Whole-body Residue concentration associated with NOEC 

after exposing populations to cadmium 

concentrations for 154 days

Marshall, 1978

Table 10. Examples of cadmium tissue residues co-occurring with the absence or presence of adverse effects.—Continued

34 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Organism

Residue 

concentration 

(mg/kg dry 

weight)

Tissue Effect Source

Tissue residues corresponding with the presence of adverse effects (lowest LOECs from each study)

Amphipod (Gammarus 

fossarum)

240–320 Whole body Reduced survival in 20-day water or dietary 

exposures1

Jarvinen and Ankley, 1999

Amphipod (Hyalella 

azteca)

0.61 Diet Severe reductions in growth (EC50) in Hyalella

fed a diet of Cd contaminated green algae. 

Corresponding Cd exposure in water 

producing the Cd in algae EC25 was about 

1.27 μg/L cadmium at hardness 125 mg/L. No 

LOEC reported

Ball and others, 2006

30 Whole body Critical body concentration resulting in 25 

percent lethality in chronic exposures, based 

on regression analysis of bioaccumulation 

against mortality from multiple tests 

Norwood and others, 2003

28–148 Whole body Increased mortality in 42-day exposures in tap 

water

Borgmann and others, 

1991

1 Whole body Increased mortality in 42-day laboratory 

exposures in reconstituted hard water 

Stanley and others, 2005

25 Whole body Increased mortality in 42-day laboratory 

exposures in treated sewage effluent 

Stanley and others, 2005

110 Whole body Increased mortality in 42-day in situ exposures 

in treated sewage effluent stream mesocosms.

Stanley and others, 2005

Atlantic salmon (Salmo 

salar

~ 1 Whole body Reduced alevin growth corresponding to a water 

LOEC of 0.78 μg/L, see table 16

Rombough and Garside, 

1982

Bull trout (Salvelinus 

confluentus)

0.9 Whole body Reduced survival Hansen and others, 2002b

Brook trout (Salvelinus 

fontinalis)

127 Gill Mean value from wild fish collected from a 

polluted river, South Fork Coeur d’Alene 

River, Idaho

Farag and others, 1998

5.1 Gill Males died during spawning Benoit and others, 1976

50 – 65 Kidney Males died during spawning Benoit and others, 1976

Fish (multiple taxa) 0.75 Whole body 5th percentile value of a large database of 

literature relating measured whole body tissue 

residues to adverse toxicological or ecological 

effects. Value over-predicted adverse effects 

to fish communities1

Dyer and others, 2000

Mayfly (Baetis 

tricaudatus)

10 Diet (periphyton) Reduced growth Irving and others, 2003

~2.2 Whole body Behavioral changes Riddell and others, 2005a

Mayfly (Stenonema sp.) 2.2 Whole body Decreased mayfly abundance in field study of 

an effluent influenced stream. Co-occurring 

water concentrations and body burdens of Ag, 

Cr, Cu and Zn were also elevated 

Birge and others, 2000

Rainbow trout 

(Oncorhynchus mykiss)

2.8 Whole body Reduced growth inferred, however description 

of growth effects was ambiguous. Exposed 

to 4.8 μg/L Cd for 30-weeks, hardness not 

reported 

Kumada and others, 1973

Table 10. Examples of cadmium tissue residues co-occurring with the absence or presence of adverse effects.—Continued

Results 35
Organism

Residue 

concentration 

(mg/kg dry 

weight)

Tissue Effect Source

Tissue residues corresponding with the presence of adverse effects (lowest LOECs from each study)—Continued

Rainbow trout 

(Oncorhynchus mykiss)

5.6 Whole body Reduced growth inferred, however description 

of growth effects was ambiguous. Exposed to 

100 mg/kg Cd in food for 12-weeks 

Kumada and others, 1973

38 Gill Reduced growth inferred, however description 

of growth effects was ambiguous. Exposed 

to 4.8 μg/L Cd for 30-weeks, hardness not 

reported. After holding in clean water for 

10-weeks, gill residue declined 94 percent

Kumada and others, 1973

80 Gill Reduced population density and biomass in 

comparison to reference streams; gill edema 

and cell degeneration noted in 96-hour 

exposure of naïve fish in the same creek 

(dissolved Cd concentrations ~ 5 μg/L at 

hardness 60-70 mg/L.

Farag and others, 2003

13.5 Gill 43 percent mortality. Gill residue resulting from 

being fed a trout chow diet spiked with 1,419 

mg/kg

Szebedinszky and others, 

2001

100 Diet (presumably 

spiked trout chow)

Reduced growth inferred, however description 

of growth effects was ambiguous. Whole 

body Cd burden declined by 95 percent after 

6-weeks of uncontaminated diet

Kumada and others, 1973

1,407 Diet (spiked trout 

chow)

8 – 43 percent mortality. This dietary exposure 

produced similar gill residues as did 2 μg/L at 

hardness 140 mg/L. Average of two tests

Szebedinszky and others, 

2001

90–114 Diet 

(live prey)

No effects on survival, slight (5–10 percent) 

reductions in growth, Exposed via live diet of 

Lumbriculus, an aquatic oligochaete that had 

been reared in Cd spiked sediments 

Erickson and others, 2003; 

Mount and others, 

2001; D.R. Mount, 

U.S. Environmental 

Protection Agency, oral 

commun, 2005

3 Diet (spiked trout 

chow)

Cortisol (an important hormone in stress 

response) response to alarm substance 

inhibited. No effect on several other 

behavioral endpoints. This dietary exposure 

produced similar whole-body residues as did 

2 μg/L at hardness 120 mg/L

Scott and others, 2003

Snail (Potamopyrgus 

antipodarum)

14.5 Whole body (without 

shell)

Significant decreases in population growth rate 

of clones from different areas of northern 

Europe or New Zealand 

Jensen and others, 2001

Southern leopard frog 

(Rana sphenocephala)

~10 Diet (periphyton) Reduced survival, although differences from 

controls (15 percent) were not statistically 

significant

James and others, 2005

Zooplankton (Daphnia 

magna)

1,180 Whole-body 21-day LC50 McCarty and Mackay, 

1993

Zooplankton (Daphnia 

galeata mendotae)

42.8 Whole-body Residue concentration associated with LOEC 

after exposing populations to cadmium 

concentrations for 154 days

Marshall, 1978

1

Converted from wet weight to dry weight, assuming 80 percent moisture content (dry weight concentration = wet weight concentration divided by 0.2).

Table 10. Examples of cadmium tissue residues co-occurring with the absence or presence of adverse effects.—Continued

36 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Invertebrates are important food items for many fish 

species, such as salmonids, sculpin, and dace. Cadmium 

concentrations in invertebrates were measured in several 

field surveys. Surveys from locations where cadmium 

concentrations in water and sediment were greater 

than background concentrations usually resulted in 

elevated concentrations in invertebrate tissues. Cadmium 

concentrations in tissues (whole organism) of aquatic insects at 

these locations ranged from about 5 to 50 mg/kg dw (Kiffney 

and Clements, 1993; Ingersoll and others, 1994; Farag and 

others, 1998; Besser and others, 2001; Maret and others, 

2003). In this range of environmentally realistic concentrations 

of dietary cadmium, accumulation, but no overt adverse 

effects were demonstrated through studies in which cadmiumspiked food was fed to rainbow trout (table 10). In contrast, 

dietary cadmium toxicity tests with fish required greater 

than 90 mg/kg dw to even slightly reduce survival or growth. 

High mortalities were only observed at dietary concentrations 

greater than about 500 mg/kg (table 10). Reduced growth and 

survival rates were observed in studies with fish that were 

fed invertebrate diets collected from the wild that contained 

mixtures of elevated arsenic, cadmium, copper, lead, and zinc 

(for example, Woodward and others, 1994; Farag and others, 

1999); however, subsequent studies implicated arsenic as the 

likely cause (Erickson and others, 2003; Hansen and others, 

2004). Thus, the available information indicates that clearly 

adverse effects in fish such as high mortality or stunted growth 

seem to require environmentally unrealistic doses of cadmium 

in the diet. 

Aquatic Invertebrates

As with fish, tissue concentrations associated with 

adverse effects in aquatic invertebrates overlapped those 

concentrations not associated with adverse effects. With the 

amphipod Hyalella azteca, the lowest whole-body tissue 

burden associated with adverse effects was 28 mg/kg; the 

highest whole-body tissue burden with no apparent adverse 

effects was 62 mg/kg. In mayflies collected from populations 

without obvious adverse effects, whole-body tissue residues 

of as much as about 35 mg/kg cadmium were measured. In 

contrast, mayflies collected from a stream that at a location 

where the benthic community was altered and was influenced 

by a mixture of metals in effluents had average cadmium 

whole-body tissue residues of only 2.2 mg/kg. Whether 

this 10 times overlap factor truly reflects species sensitivity 

differences or simply reflects the difficulty detecting effects 

and attributing causality to observed effects in field studies is 

unknown. Possibly, undetected adverse effects were present 

at the “no-effect” locations where mayflies with high tissue 

residues were collected, and at the location with apparent 

adverse effects and low tissue residues in mayflies, the 

adverse effects may have been influenced by factors other than 

cadmium concentrations in tissues. 

Contradictory results have been reported from laboratory 

exposures with Hyalella azteca. Borgmann and others (1991, 

2004) found that the chronic toxicity of cadmium to Hyalella 

azteca under differing water-quality conditions was more 

constant if toxicity was expressed as a function of cadmium 

tissue residues, rather than the cadmium concentrations 

measured in the water. In these studies, there was only a factor 

of 2 difference between the Hyalella azteca highest no-effect 

and lowest low-effect tissue concentrations. In contrast, 

Stanley and others (2005) reported greater than a factor of 100 

difference between low-effects thresholds in three tests using 

differing water-quality conditions (table 10). 

The high variability between whole-body cadmium 

tissue residues and the presence or absence of apparent effects 

is consistent with theoretical explanations of the biological 

significance of metals in tissue residues. Rainbow (2002) 

argued that the biological significance of tissue residues of 

cadmium (and other trace metals) varies greatly depending 

on the specific organisms and tissues involved. All aquatic 

invertebrates take up and accumulate trace metals, whether 

or not they are essential such as copper and zinc or nonessential such as cadmium. However, the significance of an 

accumulated concentration depends on the specific tissue 

and invertebrate involved and whether an accumulated trace 

metal concentration is high or low cannot be assessed on an 

absolute scale. Even in the absence of anthropogenic metal 

contamination and limited comparisons within the single taxon 

Crustacea, mean whole body cadmium tissue concentrations 

vary by at least a factor of 150, ranging from less than 0.3 to 

53 mg/kg dw. Accumulated metal concentrations in aquatic 

invertebrates can be interpreted in terms of different trace 

metal accumulation patterns, dividing accumulated metals 

into two components—metabolically available metal and 

stored detoxified metal. Toxicity is related to a threshold 

concentration of metabolically available metal and not to total 

accumulated metal concentration. Once an invertebrate passes 

a threshold concentration of metabolically available metal, the 

invertebrate will suffer toxic effects, initially sublethal, but 

eventually lethal. In contrast, the amount of detoxified metal 

that an invertebrate can accumulate is theoretically unlimited 

(Rainbow, 2002).

In most short-term “chronic” laboratory toxicity 

tests, during most exposures (which often are much higher 

than most environmental availabilities) much or all of the 

incoming metal remains in the metabolically available pool 

and death will occur at about the same accumulated total 

body concentration, particularly if only a small proportion of 

total body metal is held in detoxified form. Rainbow (2002) 

suggested that the apparent relation between total body 

concentration and toxicity is a consequence of experimental 

design seeking toxic effects in the short-term, as opposed to 

field situations with indefinite exposure durations. In field 

situations, most accumulated metal might be expected in the 

detoxified state, which would contribute to the poor relations 

sometimes encountered between toxic effects and whole 

body tissue residues of trace metals in aquatic invertebrates 

(Rainbow, 2002).

Results 37
Dietary cadmium exposures appear to be an important 

risk for at least some invertebrates. The data reviewed on 

dietary effects of cadmium to invertebrates indicated that 

adverse effects could occur at concentrations realistic in 

cadmium-polluted waters. Toxicity to mayflies from feeding 

on cadmium-contaminated algal mats at environmentally 

realistic concentrations was observed (Irving and others, 

2003). Although short-term acute testing with the same 

mayfly species provided acute toxicity values far higher 

than environmentally relevant concentrations, no chronic 

water-only tests are known, making direct comparisons of 

dietary versus waterborne exposures impossible. In studies 

of the relative importance of water and food as cadmium 

uptake sources and toxicity, cadmium was more toxic to the 

freshwater crustacean Daphnia magna in water-only exposures 

than in combined cadmium exposures through both water and 

diet (Barata and others, 2002b). Barata and others (2002b) 

results do not appear to support generalizations about the 

additive effects of combined cadmium exposures via food and 

water.

A diet of cadmium-contaminated green algae 

Chlorella sp caused reduced growth in the amphipod 

Hyalella azteca in a recent study (Ball and others, 2006). 

In this study, pronounced growth reductions resulted from 

food contaminated with cadmium at the lowest residue 

concentrations reviewed (0.3 to 0.6 mg/kg dry weight, table 

10). The cadmium concentrations in the medium used to 

produce the contaminated algae for the feeding tests were 

also quite low, about 0.57 and 1.27 μg/L for the contaminated 

diet-based EC25 and EC50 values respectively (table 10). 

These water concentrations overlap the chronic criterion value 

of 0.65 μg/L dissolved cadmium calculated at the culture 

hardness of 125 mg/L calcium carbonate. In this study, little 

or no bioaccumulation of cadmium was found in the tissues of 

Hyalella azteca that were fed contaminated food. These results 

demonstrate an apparent indirect or direct toxicological effect 

of cadmium-contaminated algae to Hyalella azteca that is not 

associated with cadmium accumulation. Indirect effects could 

include behavioral responses (for example, the Hyalella azteca

detected something in the contaminated food and ate less) 

or decreased nutritional quality of the food. Possible direct 

effects could result from a negative effect on the gut lining and 

thus, a reduced ability to assimilate nutrients (Ball and others, 

2006).

Bioconcentration and Bioaccumulation Factors

A simple approach for estimating tissue reference 

values (TRV) for cadmium that might result from exposure 

to waterborne exposure is to multiply the waterborne 

concentration of interest (for example, criteria value or 

concentration at a site) by bioconcentration (BCF) or 

bioaccumulation factors (BAF). When used with a chronic 

criteria value, the TRV may be further interpreted as a tissue 

residue in aquatic biota above which adverse ecological 

effects may occur (Shepard, 1998; Dyer and others, 2000). For 

example, a chronic cadmium criterion (CCC) value multiplied 

by a mean BCF for cadmium yields a tissue reference value of 

about 1 mg/kg dw for aquatic organisms as: 

TRV = CCC  ́ BCF = 0.41 μg/L  ́ 2,623 L/kg tissue as 

wet weight = 1,075 μg/kg = 1.08 mg/kg ww.

The BCF is a ratio of internal concentrations of a 

contaminant in the organs or whole body of an organism 

to the concentrations in water (McGeer and others, 2003). 

Tissue residues are reported in the literature either as dry 

weight or as wet or fresh weight. Dry weight residues reflect 

the concentrations remaining in tissue samples after water 

is removed by heating or freeze drying, whereas wet or 

fresh weight concentrations are calculated and reported on 

their concentration in the original tissue matrix. Dry weight 

normalized concentrations may be more comparable across 

studies because they correct for the different water content of 

diets or tissues, which in studies listed in table 10 ranged from 

about 20 percent in trout chow to 90 percent in zooplankton. 

In contrast, wet or fresh weight tissue residues are more 

relevant for evaluating ecological risk because no animals dry 

their prey before eating. Because most of the data reviewed 

were reported on a dry weight basis, they are interpreted here 

on a dry weight basis. Thus for use with dry weight data, 

the TRV equation needs an adjustment for water content. 

Assuming a typical water content of 80 percent, the wet 

weight tissue TRV from the above example is multiplied by 

5, yielding a dry weight cadmium TRV of 5.4 mg/kg dw (for 

example, cadmium TRV = 1.08 mg/kg ww ≈ 5.4 mg/kg dw).

Cadmium TRVs were calculated in this manner for a 

range of BCF estimates and chronic criteria values (table 10). 

These calculated values provide crude benchmarks to compare 

against measured concentrations of cadmium in tissues or 

diets that co-occurred either with adverse effects or no obvious 

adverse effects to various organisms. These comparisons 

in give some indication of whether continuous exposure to 

the water-based chronic cadmium concentration would be 

protective of from adverse effects via dietary bioaccumulation. 

TRVs ranged from 0.5 to 14.6 mg/kg dw when calculated 

over a range of BCF estimates and over a range of hardnessdependent criterion values. Qualifications such as “crude” and 

“some” are needed because of the significant uncertainties 

in calculating and interpreting cadmium TRVs calculated 

in this manner. McGeer and others (2003) found that 

BCFs for cadmium and other metals were characterized 

by extreme variability in mean values and a clear inverse 

relationship between BCF/BAF and aqueous exposure. The 

highest BCF values were at low and naturally occurring 

exposure concentrations. The inverse BCF to exposure 

relation illustrates that, although cadmium concentration 

increases with exposure, internal accumulation does not 

rise as quickly as exposure levels and therefore indicates a 

significant degree of control over cadmium accumulation. 

McGeer and others (2003) found that if BCFs were limited 

to the cadmium exposure range of 0.1 to 3 μg/L, the range 

where chronic effects were predicted to begin, mean BCFs 

38 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
were higher than BCFs calculated for all data (2,623 versus 

1,866, with coefficients of variability of 230 and 260 percent, 

respectively). They also derived modified accumulation 

factors that accounted for only the increased accumulation 

that arises from an increase in exposure concentration by 

removing the preexisting concentrations (from controls) from 

the calculation. Their mean modified accumulation factor 

was considerably lower than the conventionally calculated 

BCFs (352 versus 1,866). McGeer and others (2003) 

cautioned that using the weight of evidence available, it was 

virtually impossible to derive a meaningful BCF value that is 

representative of the BCF for each metal. Even when BCFs 

are limited to the exposure range where chronic toxicity might 

be expected (based on water-quality guidelines), it was not 

possible to derive a precise and accurate BCF value (McGeer 

and others, 2003). Nonetheless, if McGeer and others 

(2003) modified accumulation factor was considered the 

best available estimate of a typical BCF, when used with the 

CCC for a moderate hardness of 50 mg/L, the resulting TRV 

(0.41 μg/L  ́ 352 L/kg  ́ 5 = 0.7 mg/kg dw ) is lower than 

the lowest adverse effect associated with a dietary exposure or 

tissue residue in table 10. 

These calculations used BCFs that were usually derived 

from laboratory exposures where the only route of exposure 

to elevated concentrations is through the water. When derived 

from field studies, the term bioaccumulation factors (BAFs) 

is used to indicate that the exposure routes include both food 

and water. BAFs are calculated the same way as BCFs, but 

in practice reliable BAFs are more difficult to determine than 

BCFs because under field conditions it is difficult to make 

enough measurements of the material in water to show it was 

reasonably constant for long enough period of time over the 

territory range of the organisms (Stephan and others, 1985). In 

a cursory comparison of cadmium BAFs to BCFs, BAFs based 

on mean water concentrations and whole body invertebrate 

tissues (mayflies and amphipods) were calculated from 

13 exposures reported in five studies (Farag and others, 1998; 

Birge and others, 2000; Besser and others, 2001; Riddell and 

others, 2005a; Stanley and others, 2005). BAFs, calculated on 

a wet weight basis to be comparable to reported BCFs, ranged 

from 88 to 4,260 with a mean of 1,389 and a coefficient of 

variation of 101 percent. This mean BAF value is intermediate 

to McGeer and others’ (2003) mean values, which ranged 

from 352 to 2,623 depending on the method calculated. 

In summary, the data reviewed on effects of cadmium 

tissue-residues in fish and invertebrates were insufficient 

to analyze quantitatively similarly to data on the effects of 

waterborne cadmium. Qualitatively, the data reviewed suggest 

that fish can tolerate cadmium gill and kidney tissue residues 

as much as at least 4 and 20 mg/kg dw, respectively, without 

obvious adverse effects. Cadmium in the diet apparently 

poses little risk to rainbow trout in concentrations as much as 

about 50 mg/kg dw. However, although predatory fish such 

as rainbow trout may be able to tolerate feeding on prey with 

tissue residues greater than about 50 mg/kg dw, these tissue 

residues may be harmful to the prey species themselves. Some 

adverse effects with mayflies were associated with wholebody cadmium tissue residues as low as 2 mg/kg dw, however 

other data suggested some mayflies can tolerate cadmium 

tissue residues as much as about 30 mg/kg dw (table 10). 

Estimated tissue residues that might result from continuous 

exposure to chronic cadmium criteria concentrations were 

near the low range of residue concentrations associated with 

adverse effects.

In summary, the data reviewed on bioaccumulation 

and effects of dietary exposures to cadmium indicate that 

at chronic criterion concentrations, cadmium is unlikely to 

bioaccumulate to tissue residue levels expected to cause 

obvious adverse effects to aquatic invertebrates or fish. 

However, the data reviewed were not sufficient to define 

quantitatively a maximum acceptable dietary intake or tissue 

concentration for fish or other aquatic-dependent life based 

on observations of survival, growth, or reproduction in a 

chronic feeding study or a long-term field study. Therefore, 

no “final residue value” criterion based upon tissue residues 

and bioaccumulation or bioconcentration factors could be 

appropriately derived as described by Stephan and others 

(1985).

Behavioral Toxicity

Cadmium has been shown to cause neurotoxic effects 

in fish. These neurotoxic effects may manifest themselves 

through altered behavior, which in turn may predict more 

serious effects including reduced growth, reproductive failure, 

and death. Hyperactivity probably is the most widely observed 

maladaptive behavior reported from cadmium exposed fish, 

with several reports involving a variety of fish species during 

long-term cadmium exposures. Most fish that exhibited 

hyperactive behavior in long-term exposures ultimately died 

(Eaton, 1974; Benoit and others, 1976; Spehar and others, 

1978b; Sullivan and others, 1978; McNicol and Scherer, 1991; 

Riddell and others, 2005a, 2005b). In contrast to this pattern, 

Schreck and Lorz (1978) noted that cadmium exposed coho 

salmon never behaved violently, but appeared sedated up until 

the time of death with no increase in serum cortisol levels, 

a stress indicator. Hyperactivity is detrimental to small fish 

because it makes them more likely to be seen and attacked 

by predatory fish. Similarly, hyperactive predatory fish have 

lower success rates in detecting, orienting to, attacking, and 

swallowing prey (Sullivan and others, 1978; Kislalioglu and 

others, 1996; Riddell and others, 2005b). 

The mechanisms and sites of action of neurotoxic effects 

are unclear, because unlike metals such as aluminum, copper, 

lead, and mercury, cadmium apparently does not pass the 

blood-brain barrier, does not accumulate in the higher centers 

of the brain, and does not alter brain neurochemical activity 

(Beauvais and others, 2001; Scott and others, 2003; Scott and 

Results 39
Sloman, 2004). Several studies have found that cadmium did 

not accumulate in brain tissues (Kumada and others, 1973; 

Benoit and others, 1976; Thomas and others, 1983; Brown 

and others, 1986). Scott and others (2003) observed that 

cadmium could not be detected in the brain tissues of rainbow 

trout that showed maladaptive behavior following cadmium 

exposure. However, cadmium deposition in the olfactory 

system (rosette, nerve, and bulb) during waterborne exposure 

was greater than in all other organs of accumulation except 

the gill. No cadmium deposition was detected in the lateral 

line of the trout, despite that the lateral line has neural cells 

that are in direct contact with water. Lateral line function and 

rheotaxis was inhibited in a New Zealand galaxid following 

exposure to 2 μg/L cadmium. Rheotaxis, the behavior of fish 

to turn to face a current enabling them to hold position in 

moving water, depends upon lateral line function (Baker and 

Montgomery, 2001). Galaxids are stream-resident fish that 

hold a similar ecological niche to that of salmonids in the 

northern hemisphere. 

Preference or avoidance testing with fish also has 

been reported for cadmium. Fish were given the choice of 

selecting parts of a test chamber with or without cadmium. 

However, of the types of behavioral testing reviewed, 

preference/avoidance testing may be the least relevant to 

natural waters. In natural waters fish likely select and move 

among habitats based on myriad reasons such as access to 

prey, shelter from predators, shade, velocity, temperature, 

and interactions with other fish. In contrast, laboratory 

preference or avoidance tests are commonly conducted under 

simple, highly artificial conditions to eliminate or minimize 

confounding variables other than the water characteristic 

of interest. Such tests may overestimate the protection this 

behavior provides fish in heterogeneous, natural environments. 

Using lake whitefish (Coregonus clupeaformis), R.E. McNicol 

and colleagues at the Freshwater Research Institute, Winnipeg, 

extensively investigated the interactions between preference 

and avoidance of cadmium and environmental factors such 

as light and temperature. Whitefish could detect cadmium 

concentrations in a counter-current trough as low as 0.2 μg/L. 

Under uniform test conditions, fish displayed a bimodal 

concentration-avoidance response relation. Fish generally 

avoided concentrations less than or equal to 5 μg/L and greater 

than 50 μg/L, but showed little response to concentrations in 

between. However, when cadmium was added to a shaded area 

of the trough, attraction to the shade completely suppressed 

cadmium avoidance for concentrations as much as 10 times 

those avoided under homogeneous lighting. Similarly, when 

cadmium was added to a thermally attractive region of the 

trough, avoidance of cadmium was again suppressed, with the 

degree of suppression varying with the strength of temperature 

attraction. Mildly attractive 15°C water completely suppressed 

cadmium avoidance up to a concentration of 50 μg/L. When 

cadmium was presented opposite lethal temperature water 

(24°C), fish displayed a strong aversion to the warm water 

which suppressed cadmium avoidance at all test concentrations 

(McNicol and Scherer, 1991; McNicol, 1997; McNicol and 

others, 1999). Thus, when the attractive influence of a natural 

factor opposed the repelling influence of a chemical stressor, 

avoidance of the chemical stressor was suppressed. It follows 

that results of simple laboratory preference/avoidance testing 

have little relevance in complex natural habitats.

Cadmium exposures have also been shown to reduce the 

ability of fish to respond to natural chemical attack alarms. 

After a predator breaks the skin of a fish, the fish releases a 

chemical alarm signal from fish skin epithelial cells. Scott 

and others (2003) found that waterborne exposure to 2 μg/L 

cadmium for 7 days eliminated the normal antipredator 

behaviors exhibited by rainbow trout in response to alarm 

substance, whereas exposures of shorter duration or lower 

concentration had no effect on normal behavior. Predator 

avoidance behavior is complimentary to predatory behavior. 

Riddell and others, (2005a and 2005b) observed that brook 

trout exposed to sublethal cadmium concentrations had 

reduced success in capturing their preferred mayfly prey, 

subsequently switching their prey preferences to less motile 

midges, with consequences of reduced growth. The LOEC 

for the behavioral alterations of 0.4 μg/L at a hardness of 

156 mg/L was lower than chronic test endpoints for brook 

trout.

Other fish behaviors that may be disrupted by cadmium 

exposures include social behavior, dominance hierarchies, and 

prey selection (Sloman and others, 2003a, 2003b, and 2005; 

Scott and Sloman, 2004). Long term exposure of rainbow 

trout to 3 μg/L cadmium in water with a hardness of 140 mg/L 

had no effect on their critical swimming speed (McGeer and 

others, 2000a).

In summary, cadmium exposures may result in neurotoxic 

effects that may manifest themselves through altered behavior, 

which in turn may predict more serious effects including 

reduced growth, reproductive failure, and death. Generally, 

behavioral responses were observed at lower cadmium 

concentrations than concentrations causing mortality. Complex 

behaviors provide the foundation for fish population structure 

and aquatic communities, and if natural behaviors were 

fundamentally altered by cadmium or other pollutants, that 

would have serious implications for aquatic ecosystems. 

40 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Sensitivity of Different Life Stages of Aquatic Organisms 

to Cadmium

The data reviewed generally supported (or at least did 

not clearly refute) the assumption that juvenile fish and early 

instars of invertebrates were more sensitive to cadmium 

than older life stages. The most information was available 

for fish (salmonids, sculpin, and fathead minnow) and two 

commonly tested invertebrates, Daphnia magna and Hyalella 

azteca. However, the data reviewed on size or age relations 

to cadmium sensitivity were not always pronounced or 

consistent. 

Results of experiments that have been reported for 

different life stages or ages of freshwater aquatic invertebrates 

commonly used for toxicity testing were equivocal regarding 

identification of any particular life stage that is consistently 

most sensitive to metals. For example, sensitivity of 

Chironomus tentans to cadmium decreased in advanced 

developmental instars of the midge, with the first instar 

being more than 100 times more sensitive than later instars 

(Williams and others, 1986). Conversely, the sensitivity of 

Daphnia magna to cadmium and copper was similar for 

organisms ranging from less than 1 to 6 days old (Nebeker 

and others, 1986a). Collyard and others (1994) found that 

the sensitivity of Hyalella azteca to toxicants in short-term 

exposures does not differ greatly as a function of age over a 

range of <1–26 days old at test initiation. In the latter tests, 

96-hour LC50 values for cadmium chloride varied by about a 

factor of 2 among six age classes of Hyalella azteca ranging 

from 0 to 2 to 24 to 26 days. Organisms 10–12 days old at test 

initiation had the lowest cadmium LC50, however, differences 

between the age groups were small (Collyard and others, 

1994). Wigginton (2005) reported that the early life stages of 

species of crayfish in the family Cambaridae were 10 to 40 

times more sensitive to cadmium than adults.

Among several life stages of Chinook salmon and 

steelhead (that is, the anadromous form of the rainbow 

trout), Chapman (1978b) found that with Chinook salmon 

the swim-up fry was the most sensitive life stage, but with 

steelhead, swim-up fry and parr life stages were similarly 

sensitive to acute-cadmium exposures, whereas both younger 

(alevin) and older and larger (smolt and adult) life stages 

were considerably more resistant.2 In contrast, steelhead 

exposed to copper and zinc at the swim-up stage were 

more sensitive than parr. Chapman (1975) found a slightly 

different pattern with coho salmon than with Chinook and 

steelhead. Six rangefinding acute tests were conducted using 

fry hatched from the same batch of eggs to evaluate the 

sensitivity of different aged fish. Results suggested that the 

2

 Alevins are fish that have hatched but still contain a yolk sac. Swim-up 

fry are fish that are 4–6 weeks post hatching, have resorbed their yolk sacs 

and have begun feeding, and have just swum up from their incubation cups or 

natural gravels to swim freely in the water column. Parr are larger juveniles 

that may be up to two years old.

youngest swim-up fry and older juveniles were more resistant 

than 6-week old fry. Chapman and Stevens (1978) found 

that the greater resistance of adult salmonids to cadmium 

than juveniles was only temporary. When measured over 4 

days, adults were much more resistant than juveniles, but 

when measured over 17 days, the adults were as sensitive as 

juveniles.

Hansen and others (2002a) noticed different patterns of 

sensitivity of rainbow trout and bull trout within the fry life 

stage. Older and larger (about 1 g) fish were more sensitive 

than younger and smaller (about 0.1 to 0.2 g) fish. Stubblefield 

and others (1999) reported very small (0.19 g average weight) 

swim-up fry rainbow trout were less sensitive to cadmium than 

were adults (table 17).

For mottled sculpin, Besser and others (2007) found 

that newly hatched fish were much more sensitive to 

cadmium, copper or zinc than older juveniles or yearlings. 

Tests with mottled sculpin conducted with 3-week, 7-week, 

3-month old, and yearling fish showed a consistent pattern of 

increasing resistance with age. Unlike the patterns observed 

with salmonids for cadmium and copper, the newly hatched 

sculpins were even more sensitive than the swim-up fry.

Most life-stage sensitivity data reviewed was consistent 

with the notion that juveniles were probably more sensitive 

than older organisms, and salmonid fry about 6–10 weeks 

posthatch or about 1 g in weight may be particularly 

vulnerable. However, the available data were too inconsistent 

to consider a priori life stages to be sensitive or resistant. 

Thus, if data for only one life stage were available, it was 

considered usable in table 15. If data for more than one life 

stage were available, they were used in table 15 or rejected 

on a case-by-case basis. Data rejected because tests used 

apparently resistant life stages included: (1) all acute results 

for adult fathead minnows because they were consistently 

more resistant than fry; (2) results of a flow-through acute test 

with yearling brook trout that were about 1,000 times higher 

than an acute test with brook trout fry; and (3) results of a 

flow-through test with 6-month old coho salmon (table 17).

Acclimation

Exposure to sublethal concentrations of cadmium and 

other metals may result in pronounced increased resistance to 

later exposures of the organisms. Several studies concluded 

that acclimation is a real compensatory response that plays 

an important role in the existence of natural fish populations 

in metals polluted environments. However, the increased 

resistance of fish to metals exposure can be temporary and 

lost in as few as 7 days after return to unpolluted waters 

(Bradley and others, 1985; Sprague, 1985; Hollis and others, 

1999; Stubblefield and others, 1999). For this reason, EPA’s 

guidelines for deriving aquatic life criteria specify that test 

results from organisms that were pre-exposed to toxicants 

should not be used in criteria derivation (Stephan and others, 

1985). 

Results 41
Screening results of short-term acute tests for preexposure to toxicants is reasonably straightforward. Evaluating 

the role of acclimation in longer term, chronic tests is less 

so. Several tests have shown that at least with fish, if toxicity 

tests were initiated during metals resistant life stages (adult 

or embryo), acclimation may occur. Later in the test when the 

more sensitive life stages become exposed (fry or juvenile), 

usually sensitive life stages may be more resistant than the 

same life stages of naïve fish which had no pre-exposure 

(Chapman, 1978a, 1994; Spehar and others, 1978b; Farag and 

others, 1994; Brinkman and Hansen, 2007; De Schamphelaere 

and Janssen, 2004). 

Chapman (1978a) exposed adult sockeye salmon 

(Oncorhynchus nerka) to a series of zinc concentrations for 

3 months followed by an 18-month exposure of embryonic 

through smolt stages. When 9-month old parr salmon were 

tested for acute resistance to zinc, the acclimated fish were 

2.2 times more resistant than the control fish. In further 

experiments, Chapman (1994) exposed different life stages 

of steelhead (Oncorhynchus mykiss) for the same duration 

(3 months) to the same concentration of copper (13.4 μg/L at 

a hardness of 24 mg/L as CaCO3). The survival of steelhead 

that were initially exposed as embryos was no different than 

that of the unexposed control fish, even though the embryos 

developed into the usually sensitive swim-up fry stage during 

the exposure. In contrast, steelhead that were initially exposed 

as swim-up fry without the opportunity for acclimation during 

the embryo state suffered complete mortality.

Brinkman and Hansen (2007) compared the responses of 

brown trout (Salmo trutta) to long-term cadmium exposures 

that were initiated either at the embryo stage (early-life stage 

tests) or the swim-up fry stage (chronic growth and survival 

tests). In three comparative tests, fish that were initially 

exposed at the swim-up fry stage were consistently 2–3 times 

less resistant than were the fish initially exposed at the embryo 

stage.

These studies support the counterintuitive conclusion 

that because of acclimation, longer term tests or tests that 

expose fish over their full life cycle are not necessarily more 

sensitive than shorter term tests which are initiated at the 

sensitive fry stage. This conclusion has important implications 

for relating test results to potential effects in the wild. Some 

groups of temperate freshwater fishes, such as salmonids and 

suckers, have migratory life histories. Their life histories often 

involve spawning migrations to headwater reaches of streams, 

followed by downstream movements of fry shortly after 

emerging from the substrates, and later seasonal movements 

to larger, downstream waters to over winter (Willson, 1997; 

Baxter, 2002; Cooke and others, 2005; Quinn, 2005). These 

life history patterns often correspond to human development 

and metals pollution patterns such that headwater reaches 

likely have the lowest metals concentrations and downstream 

increases could occur due to point source discharges or 

urbanization. At least for salmonids and suckers, the scenario 

of metals acclimation developing at the egg stage in upstream 

spawning reaches that would provide fry and older juvenile 

stages additional protection seems unlikely. Therefore, full life 

cycle toxicity testing or early life stage testing methods may 

not necessarily provide “safer” or more relevant estimates of 

acceptable long-term exposure concentrations than shorter 

term methods initiated at sensitive life stages.

Evaluation of Risks to Biological Integrity and 

Vulnerable Species

The preceding descriptions of effects of cadmium on 

species and criteria derivation focused on the results of testing 

single-species with cadmium under constant conditions in 

laboratory settings. This reductionist approach to criteria 

derivation was needed to eliminate confounding variables and 

to make comparisons of species-sensitivities internally valid. 

This of course is not the case in the real world where species 

interact and contaminants occur in mixtures and in variable 

concentrations. 

The criteria values derived in this report were evaluated 

in the context of whether waters with cadmium concentrations 

near criteria conditions would be expected to maintain 

biological integrity. This effort included: 

1. Evaluating the criteria concentrations that were 

derived from toxicity tests conducted in laboratories 

under constant conditions in the context of 

fluctuating cadmium concentrations in field 

conditions; 

2. The assumption that the most sensitive 5 percent 

of taxa from a species-sensitivity distribution need 

not be explicitly protected in criteria derivation was 

critically examined by considering the role that a 

more-sensitive “unprotected” species has in aquatic 

food webs and through population modeling; and

3. Comparing apparent effects concentrations from 

field surveys or ecosystem studies to corresponding 

criteria concentrations. 

Criteria in Variable Environments

One major problem with estimating chemical criteria that 

are neither underprotective nor overprotective is how to bridge 

the gap between the nearly constant concentrations used in 

the toxicity tests used to derive the criteria concentrations 

and the fluctuating concentrations that usually exist in the 

real world (Stephan and others, 1985). Stephan and others 

(1985) addressed the problem by defining averaging periods 

for criteria. The intent of their criteria guidelines was to 

avoid “unacceptable” effects, and some adverse effects such 

as a small reduction in the survival, growth, or reproduction 

42 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
even in important species was considered possible at or 

less than criteria. The criterion continuous concentration 

(CCC) was intended to be a good estimate of this threshold 

of unacceptable effect. If maintained continuously, any 

concentration greater than the CCC was expected to cause 

an unacceptable effect. However, they expected that the 

concentration of a pollutant in a water body could be greater 

than the CCC without causing an unacceptable effect if (1) 

the magnitudes and durations of the excursions were limited 

and (2) if there were compensating periods of time in which 

the concentrations were less than the CCC. The more the 

concentration was greater than the threshold of unacceptable 

effects, the shorter the period of time that it could be tolerated. 

But, whether there was any upper limit on the magnitude 

of exceedences that could be tolerated for very short period 

of times, such as on the order of 1 minute, was considered 

unimportant because concentrations in ambient waters rarely 

change substantially in such short periods of time (Stephan 

and others, 1985). 

Stephan and others (1985) considered an averaging 

period of 1 hour most appropriate to use with the criterion 

maximum concentration (CMC) because high concentrations 

of some materials could cause death in 1 to 3 hours. 

Additionally, even when organisms do not die within the 

first few hours, few toxicity tests continue to monitor for 

delayed mortality after the exposure period ends. Thus it 

was considered inappropriate to allow concentrations greater 

than the CMC for more than 1 hour (Stephan and others, 

1985). U.S. Environmental Protection Agency (2001) used a 

24-hour averaging period for their cadmium CMC, although 

no explanation was given for the departure from Stephan and 

others (1985) guidelines.

Stephan and others (1985) guidelines on this point were 

based upon their collective experience and judgment rather 

than specific datasets. However, subsequent research on 

delayed toxicity in pulsed exposures and relations between 

metals accumulation on fish gill surfaces and toxicity is 

relevant to the selection of criteria averaging periods. Marr 

and others (1995) episodically or continuously exposed 

fish to a metals mixture; the concentrations and ratios of 

the metals and variations in water quality (pH, hardness) of 

which were selected to represent conditions measured during 

episodic storm events in the Clark Fork River, Montana. 

Mortality resulted from 8-hour exposures followed by 96-hour 

observation periods. At decreased hardness and constant pH 

(simulating hardness decreasing with runoff), the 8-hour LC50

was 4 times that of a 96-hour LC50 (Marr and others, 1995). In 

pulsed exposure chronic toxicity tests with fathead minnows, 

a single 6-hour pulse of cadmium at 3 times the 7-day chronic 

value followed for 7-days observation without further metals 

exposure resulted in slightly (5 percent) reduced survival 

(Where the 7-day chronic value was 13 μg/L (EC25) and the 

pulse was 40 μg/L, tests were at a hardness of 90–110 mg/L). 

A 12-hour pulse followed for 7 days resulted in high mortality 

(61 percent). A single 5 times pulse for 6 hours (60 μg/L) 

resulted in very high mortality at the end of 7 days (Diamond 

and others, 2005). In other tests, amphipods (Hyalella azteca) 

were exposed to different toxicant concentrations, and the 

percentage of mortalities was noted both during and after the 

exposure ended. For copper at the conventional 48-hour LC50

concentrations, the predicted proportions dead after including 

latent mortality were 65 to 85 percent, not 50 percent. Because 

of latent mortality, “true” Hyalella azteca LC50s with copper 

could be about one-half those calculated immediately at the 

end of acute exposures (Zhao and Newman, 2004). 

When fish are exposed to cadmium or copper, a relatively 

rapid increase occurs above background levels of metal bound 

to the gill. This rapid increase occurs on the order of less than 

3 to 24 hours, and this brief exposure has been sufficient to 

predict toxicity at 120 hours (Playle and others, 1993; Playle, 

1998; MacRae and others, 1999; Di Toro and others, 2001). 

Acute exposures of 24 hours might not result in immediate 

toxicity, but deaths could result over the next few days. Simple 

examination of the time-to-death in 48- or 96-hour exposures 

would not detect latent toxicity from early in the exposures. 

Observations or predictions of appreciable mortality resulting 

from metals exposures on the order of only 3 to 6 hours 

supports the earlier recommendations by Stephan and others 

(1985) that the appropriate averaging periods for the CMC is 

on the order of 1 hour.

Stephan and others (1985) considered a 4-day averaging 

period appropriate to use with the CCC for two reasons. First, 

“chronic” responses in some species may not really be due 

to long-term stress or accumulation, but rather the test was 

simply long enough that a briefly occurring sensitive stage 

of development was included in the exposure (Chapman, 

1978a; Barata and Baird, 2000). Second, a much longer 

averaging period, such as 1 month would allow for substantial 

fluctuations above the CCC. Substantial fluctuations may 

result in increased adverse effects from those expected in 

constant exposures (Stephan and others, 1985). A comparison 

of the effects of the same average concentrations of copper 

on developing steelhead, Oncorhynchus mykiss, that were 

exposed either through constant or fluctuating concentrations 

found that steelhead were about twice as resistant to 

constant exposures as to fluctuating exposures. In daily 

fluctuating copper exposures that peaked at the 96-hour LC50

concentration for 7 hours, the onset of growth reduction 

occurred at an average copper concentration of 16 μg/L 

copper, whereas in constant exposures, the onset of growth 

reduction occurred at an average concentration of 30 μg/L 

(Seim and others, 1984). A similar test design using cadmium 

found no increase in chronic toxicity resulting from fluctuating 

exposures of Daphnia pulex over constant exposures. In this 

Results 43
case, the Daphnia were only exposed daily for 70 minutes 

to cadmium pulses at their 48-hour LC50 concentration that 

produced a daily-average cadmium exposure that was similar 

to the chronic cadmium criterion concentration in effect at the 

time (Ingersoll and Winner, 1982). These latter results indicate 

that fairly high cadmium exposures of about 1 hour may be 

tolerated by some organisms.

Measured cadmium fluctuations in several regional 

streams were examined to estimate how ambient fluctuations 

compared to experimental toxicity results from pulsed 

exposures to cadmium (fig. 3). Comparisons were limited to 

waters for which multiple cadmium detections and matched 

calcium and magnesium measurements were available (the 

latter were needed to calculate hardness and chronic criterion 

values). Differences between average long-term concentrations 

and peak-measured concentrations ranged from 1.1 to 5 times 

with the median average-to-peak difference for all sites of 

2.2 times, with the exception of a single anomalous value from 

Thompson Creek that was 30 times higher than the median 

value for that stream. These comparisons suggest that if the 

results of the experiment with fluctuating copper exposures 

and steelhead are relevant to cadmium, then the potential 

increased sensitivity of about 2 times caused by fluctuating 

concentrations would be roughly offset by the regional pattern 

of peak cadmium concentrations being about 2 times higher 

than median concentrations.

In addition to the seasonal patterns shown in figure 3, 

daily fluctuations can be important in some streams. In 

streams draining historical mining areas in Idaho and 

Montana, cadmium concentrations in streams increased greater 

than 2 times from afternoon minimum values to maximum 

values shortly after sunrise (Nimick and others, 2003; Nimick 

and others, 2005). These patterns have important implications 

for water-quality monitoring because people are more apt to 

collect water samples in the afternoon than at sunrise. Because 

similar relations were observed in streams in which metals 

concentrations were greatly elevated above background as 

in streams in which metals are closer to background levels 

(Nimick and others, 2005), daily cycles may be important 

from a toxicological view.

Daily fluctuations in exposure conditions results in 

chemical non-equilibrium, which likely changes the toxicity 

of metals. However, the magnitude and even the direction 

of effect (more or less toxic) are unclear. The toxicological 

relevance of non-equilibrium exposure conditions probably 

has been studied best with copper and dissolved organic 

matter, where about 24 hours are needed for equilibrium, and 

the toxicity of copper decreased when added copper was given 

longer to equilibrate with dissolved organic matter (Kim and 

others, 1999: Ma and others, 1999). With cadmium, reaction 

kinetics with dissolved organic matter are less important 

than with copper. Still, the reaction of the added cadmium 

with inorganic materials in the water probably occurs at 

rates relevant to the daily cycles shown in figure 3. Davies 

and Brinkman (1994b) reported that in kinetic equilibrium 

experiments in hard, alkaline water, it took about 48 hours for 

cadmium activity to reach equilibrium. 

Davies and Brinkman (1994b) tested the relative 

toxicity of cadmium to rainbow trout in non-equilibrium and 

equilibrium exposures. In their non-equilibrium (unaged) 

exposure, the cadmium and hard water were mixed in a diluter 

and immediately discharged directly into the test aquaria. 

In their equilibrium (aged) exposure, the cadmium and hard 

water were allowed to mix in an intermediate aquarium for 

about 82 hours before overflowing into the test tanks. The 

total contact times (time for 95 percent water replacement) 

for the cadmium and hard water were about 28 hours in 

the unaged and 135 in the aged exposure. Their hypothesis 

was that cadmium would be more toxic in non-equilibrium 

conditions because by adding cadmium as soluble, free metal 

ions, toxicity would be overestimated in unaged hard water 

where insufficient time was allowed to establish equilibrium. 

Their results showed the opposite. Cadmium was more toxic 

(lower ECp and LOEC values) in equilibrium conditions than 

in non-equilibrium conditions. Davies and Brinkman (1994b) 

concluded that under conditions of non-equilibrium, results 

from toxicity tests will significantly underestimate toxicity 

because of analysis of metastable forms of metals which are 

not biologically available. (Metastable refers to an unstable 

and transient but relatively long-lived state of a chemical or 

physical system, such as of a supersaturated solution.)

In contrast to copper, where all data indicated toxicity 

was likely greater in non-equilibrium than equilibrium 

conditions (Kim and others, 1999: Ma and others, 1999), the 

only known cadmium study had opposite findings (Davies 

and Brinkman (1994b). Thus, daily fluctuations in exposure 

conditions (fig. 3) likely result in chemical non-equilibrium 

and influence cadmium toxicity. However, the data with 

cadmium are sparse, making it difficult to infer from the 

experimental results whether comparable patterns would be 

expected in different environmental conditions (for example, 

softer water, shorter exposure times.

44 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Figure 3. Variability in dissolved and chronic cadmium criterion concentrations in selected streams in Idaho and Montana, 

1997–2005. “USGS” followed by a number indicates USGS station numbers that were accessed at http://waterdata.usgs.gov/

nwis, April 2005 data from Thompson Creek and the Salmon River were provided by Bert Doughty, Thompson Creek Mining 

Company, Challis, Idaho, written commun., 2005; data from the Yankee Fork were provided by Bob Tridle, Hecla Mining 

Company, Coeur d’Alene, Idaho, written commun., 2005. 

id91U05_fig03a

0

0.1

0.2

0.3

0.5

0.4

0

1

2

3

5

4

1

0

2

3

4

5

0

5

10

15

20

0

5

10

15

20

Cadmium, in micrograms per liter Cadmium, in micrograms per liter Cadmium, in micrograms per liter

0

1

2

3

5

4

1996 1997 1998 1999 2000 2001 2002 2003 2004

2000 2001 2002 2003 1996 1997 1998 1999 2000 2001 2002 2003

1997 1998 1999 2000

00:0006:0012:0018:0000:00

September 2001

11 12 11 12 06:00 12:00 18:00 00:00 00:00 06:00 12:00 18:00 00:00

September 2001 06:00

12:0018:0000:00

Coeur d'Alene River near Cataldo, ID

USGS 12413500

North Fork Coeur d'Alene River near Enaville ID,

USGS 12413000

Coeur d'Alene River near Cataldo, ID

USGS 12413500, Nimick and others, 2003

South Fork Coeur d'Alene River near Pinehurst, ID

USGS 12513470, Nimick and others, 2003

South Fork Coeur d'Alene River near Pinehurst, ID

USGS 12513470

Coeur d'Alene River near Harrison ID,

USGS 12413000

Chronic criterion (μg/L Cd)

Dissolved Cd (μg/L)

Results 45
0

0.1

0.2

0.3

0.4

0.5

2000 2001 2002 2003

0

0.2

0.4

0.6

0.8

1.0

1996 1997 1998 1999 2000 2001 2002 2003

0

0.2

0.4

0.6

0.8

1.0

0

0

0.2

0.1

0.3

0.4

0.5

0

0.1

0.2

0.3

0.4

0.5

DEC JAN FEB MAR APR MAY JUN JUL AUG

2002 2003

2003 2004 2005

OCT NOV SEP 2000 2001 2002 2003

2003 2004 2005

0.2

0.4

0.6

0.8

1.0

Yankee Fork in mine effluent mixing

 zone near Bonanza, ID

Yankee Fork upstream of mine effluent

 outfall near Bonanza, ID

Boulder River near Boulder, MT

USGS 6032400

Pine Creek near Pinehurst, ID

USGS 112413445

Spokane River at Post Falls, ID

USGS 12419000

St. Joe River near Chatcolet, ID

USGS 12415140

Cadmium, in micrograms per liter Cadmium, in micrograms per liter Cadmium, in micrograms per liter

Chronic criterion (μg/L Cd)

Dissolved Cd (μg/L)

Figure 3. Continued.

46 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
id9IU05_fig03c

Cd (μg/L)

Chronic criterion (μg/L Cd)

Samples collected upstream of mine discharges (μg/L Cd)

Samples collected in between two discharge points (μg/L Cd)

Samples collected downstream of mine discharges (μg/L Cd)

Chronic criterion downstream of mine discharges (μg/L Cd)

Thompson Creek near Clayton, ID

Salmon River downstream of Thompson Creek Mine

discharges, near Clayton, ID

1998 1999 2000 2001 2002 2003 2004

1998 1999 2000 2001 2002 2003 2004

0

0.1

0.2

0.3

0.4

0.5

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

1.1

1.2

1.3

1.4

1.5

Cadmium, in micrograms per liter Cadmium, in micrograms per liter

Figure 3. Continued.

Results 47
Risks of the 5th Percentile Species-Sensitivity 

Distribution (SSD)-Based Chronic Criterion to a 

More Sensitive Species

One feature inherent to defining water-quality criteria 

on the basis of the 5th percentile of the SSD of single-species 

laboratory toxicity tests for the chemical is that by definition, 

of the species tested, species with mean response values 

lower than the percentile may not be protected. Aquatic 

ecosystems are assumed to have some functional redundancy 

and resiliency such that unless the more-sensitive species are 

judged to be “critical” species, avoiding adverse effects to 

95 percent of the aquatic species tested would be sufficient to 

protect the biological integrity of natural ecosystems (Stephan 

and others, 1985). 

The “5th percentile” approach has been criticized for 

depending on assumptions that may not be met or are untested. 

These include the assumptions that haphazard collections of 

data from single-species laboratory toxicity tests can be used 

to represent the sensitivities of natural ecosystems, whether 

any species loss from a community is acceptable, and if so, 

whether the 5th percentile of the SSD is the appropriate level 

of protection, among others (Forbes and Forbes, 1993a; 

Smith and Cairns, 1993; Calow and others, 1997; Power 

and McCarty, 1997; Newman and others, 2000; Forbes and 

Calow, 2002; Posthuma and others, 2002; Suter and others, 

2002b; Maltby and others, 2005). Often, fewer than 20 chronic 

values for different taxa are available, and all measured values 

are greater than the 5th percentile of the SSD. This makes 

evaluations of risks to more sensitive but unidentified taxa 

somewhat theoretical and abstract. However, when the chronic 

dataset used to construct the SSD exceeds 20, the question of 

protectiveness becomes more tangible. Effects of the chemical 

to known species with known life histories, roles in aquatic 

food chains, and measured sensitivities to the chemical can be 

evaluated. 

In the case of the acute criterion derivation, the “5th

percentile” approach probably would be unacceptable for use 

in Idaho. This is because of the 69 species for which species 

mean acute values (SMAVs) were available, the SMAVs for 

all species of trout native to Idaho fell below the 5th percentile 

of the SSD: cutthroat trout, rainbow trout, and bull trout 

(fig. 5 in the section “Field Validation,” table 7). Cutthroat 

trout were considered to be a “critical” species in the sense 

of Stephan and others (1994b) because it is a native fish 

of conservation concern and, as the state fish of Idaho, is a 

socially important sport fish. Rainbow trout are an important 

game fish and bull trout are listed as threatened under the 

Endangered Species Act. All three species are important upper 

trophic level carnivores in many streams and lakes. Basing 

the acute criterion on the lowest SMAV effectively set the 

acute criterion close to the 1st percentile of the SSD (that is, 

protecting nearly 99 percent of aquatic species). 

In the case of the chronic dataset, genus mean chronic 

values (GMCVs) were available for 21 genera, and only 

one GMCV and species mean chronic value (SMCV) fell 

below the 5th percentile of the SSD (for the amphipod 

Hyalella azteca). In contrast to the acute dataset, for the most 

chronically-sensitive species, Hyalella azteca, no a priori

judgment was made that Hyalella azteca was a “critical” 

species. Instead, risks of Hyalella azteca population declines 

and indirect food web consequences were evaluated.

The following describes the role of Hyalella azteca in 

aquatic food webs in the Pacific Northwest, the occurrence 

of Hyalella azteca in waters observed to exceed the chronic 

criterion, the degree to which Hyalella azteca toxicity 

thresholds are below the chronic criterion, and forecasts the 

likely consequences to Hyalella populations of setting the 

chronic criterion above their SMCV. 

Role of Hyalella azteca in Aquatic Food Chains

Taxonomically, Hyalella azteca is listed as the sole 

species in the family Hyalellidae, within the order Amphipoda 

(Integrated Taxonomic Information System, 2002). However, 

recent genetic studies argue that Hyalella azteca probably 

is a species complex, rather than a single species (Duan and 

others, 1997; Wang and others, 2004). These arguments are 

consistent with the wide geographic range of Hyalella azteca. 

The species was originally described from specimens collected 

from the Aztec region of southern Mexico, and the distribution 

of Hyalella azteca occurs from at least southern Central 

America, where the species was originally described, to the 

high Arctic.

Amphipods are benthic crustaceans that occupy an 

intermediate position in aquatic food webs between detritus 

and predators, such as salamanders and salmonids (Mathias, 

1971). Salmonids and other fish readily prey upon amphipods, 

probably consuming them in rough proportion to their 

abundance relative to other vulnerable invertebrates. In a lake 

in southern British Columbia, the fish Oncorhynchus mykiss

(rainbow trout) and Oncorhynchus nerka (sockeye or kokanee 

salmon) and the salamanders Taricha granulosa granulosa

(rough skinned newt) and Ambystoma gracile (Northwestern 

salamander) “feed to a significant degree on amphipods” 

(Mathias, 1971). In the lower Snake River in Washington 

and Idaho, amphipods contributed 2.7 and 7.9 percent 

of identifiable prey categories found in the stomachs of 

juvenile Chinook salmon and steelhead, respectively, from 

Lower Granite Reservoir (n=379 and 204, respectively). 

For steelhead, amphipods were the 5th most important prey 

category after Diptera larvae (for example, midges and flies, 

24 percent), insect exuviae (12 percent), mayfly nymphs 

(10 percent), and adult Diptera (9.1 percent). For Chinook 

salmon, amphipods were the 7th most important prey category 

after mayfly nymphs (27 percent), Diptera larvae (for example, 

midges and flies, 25 percent), adult Diptera (12 percent), ants 

48 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
(9.1 percent), insect exuviae (7.8 percent), and caddisflies (3.4 

percent) (Karchesky and Bennett, 1999). Although amphipods 

are often abundant in the tailwaters below reservoirs, this was 

not the case for the riverine Hells Canyon reach of the Snake 

River, Idaho. In intensive scuba and shallow water sampling 

of macroinvertebrates in the Hells Canyon reach of the Snake 

River, amphipods were relatively abundant (32 percent of the 

sample) in only 1 of 28 reaches sampled. Amphipods were 

as much as 5 to 10 percent of the samples in 4 reaches, and 

the remaining 23 reaches, comprised from 0 to 5 percent of 

the sampled assemblage. Amphipod taxa collected in the 

Hells Canyon reach of the Snake River were Crangonyx, 

Gammarus, and Hyalella (Myers and Foster, 2003). This 

relative scarcity suggests that amphipods unlikely are major 

components of the diets of Chinook salmon or other fish in the 

Hells Canyon reach of the Snake River.

In mountain lakes in Oregon, Strong (1972) noted 

that during early July rainbow trout stomachs were stuffed 

with Hyalella but by late July none could be found in 

trout stomachs. Amphipods other than Hyalella have been 

observed to be seasonally or periodically important in 

salmonid and sculpin diets and Hyalella could be similarly 

important. A larger amphipod, Gammarus lacustris, and 

Daphnia pulex dominated the diet of bull trout in an alpine 

lake in the Canadian Rockies where bull trout were the 

only species of fish present (Wilhelm and others, 1999). 

Similarly Gammarus sp. was the dominant food item in 

juvenile Chinook smolts in the lower Columbia River 

where Gammarus was abundant (Muir and Coley, 1996). 

The amphipod Crangyonx richmondensis was one of the 

dominant food forms for rainbow trout in a long-term study 

of the effects of pulp mill effluents on coldwater stream 

communities. In that study, experimental stream channels 

constructed along the lower Clearwater River, Lewiston, 

Idaho, were allowed to colonize naturally with invertebrates 

for several months before testing effluents. Trout stomach 

samples indicated the dominant food forms in control streams 

were Diptera (primarily Chironomidae), Amphipoda, and 

Gastropoda. The amphipods usually were 30 percent or more 

of trout food intake. Gastropods were only an important food 

source in autumn during stream studies (National Council 

of the Paper Industry for Air and Stream Improvement, 

Inc,, 1989). In response to the collapse of populations of the 

amphipod Diporeia sp., populations of two native benthivores, 

slimy sculpin (Cottus cognatus) and lake whitefish 

(Coregonus clupeaformis), suffered severe declines in eastern 

Lake Ontario, perhaps partly due to starvation, because 

Diporeia sp. was their principal prey (Owens and Dittman, 

2003). Amphipods in the genera Crangonyx, Diporeia, and

Gammarus appear to be less sensitive to cadmium than 

Hyalella (tables 7 and 17). 

Under favorable conditions, Hyalella azteca can be 

extremely abundant. In Blue Lakes Spring, a tributary to the 

Snake River near Twin Falls, Idaho, Hyalella azteca densities 

reached 29,400 per m2 (Maret and others, 2001). Although 

no dietary studies are known, at these densities Hyalella 

azteca likely are important in the diets of the sculpin and trout 

populations that are resident in this large spring. 

In summary, the available literature indicates that fish 

and salamanders will exploit Hyalella and other freshwater 

amphipod populations when they are abundant, but if 

other suitable prey are available, will readily switch when 

amphipods are scarce or absent. 

Hyalella azteca in Waters with Elevated Cadmium 

Concentrations

If Hyalella azteca are as sensitive to cadmium 

in the wild as laboratory studies indicate, then some 

correspondence between their occurrence and elevated 

cadmium concentrations would be expected. Biological and 

chemical data sets that were reasonably well matched in time 

and space were located from two river basins—the Clark 

Fork River, Montana, and the Coeur d’Alene/Spokane basin, 

Idaho. Hyalella azteca were present in the Spokane River 

at Post Falls, Idaho, in July 1999 (U.S. Geological Survey, 

2005). Cadmium concentrations at that location were elevated 

well above background levels, commonly approaching and 

slightly exceeding the cadmium chronic criterion (maximum 

of 1.1 times the chronic criterion, fig. 3). Amphipods were 

common at some subblittoral (5–10 m depth) locations in 

Coeur d’Alene Lake that either were uncontaminated by 

cadmium, zinc, and lead wastes or that had moderately 

elevated concentrations (1.1 to 2.0 times the chronic criterion 

at the time of sampling). Amphipods were not collected at 

more contaminated sites, but they also were not collected 

from the reference lake, Priest Lake, Idaho (Ruud, 1996). 

Ruud (1996) did not identify amphipods below “Amphipoda,” 

but because the only amphipod species collected by the U.S. 

Geological Survey from the outlet to Coeur d’Alene Lake (the 

Spokane River) was Hyalella azteca, it seems likely that the 

amphipods in Coeur d’Alene Lake included Hyalella azteca. 

Amphipods were collected from several locations in 

the Clark Fork River and Milltown Reservoir, Montana, in 

1991 as part of investigations of the ecological effects of 

metals contaminated sediment (Canfield and others, 1994). 

The amphipods were not identified below “Amphipoda,” but 

because the only amphipod species collected from tributaries 

to the Clark Fork River in the 1999 USGS National Water 

Quality Assessment (NAWQA) surveys was Hyalella azteca, it 

seems likely that the “Amphipoda” collected by Canfield and 

others, 1994, included Hyalella azteca. Sediments collected 

from the sites where the amphipods were collected were 

contaminated by a mixture of metals which caused sublethal 

toxic effects to Hyalella azteca in laboratory exposures, 

although the magnitude of adverse effects was fairly low 

(Canfield and others, 1994). A cadmium concentration 

of 1.1 μg/L at a hardness of 100 mg/L was considered 

representative for the upper Clark Fork River (Woodward 

and others, 1994), which would be 1.75 times the chronic 

Results 49
criterion derived in this report. This suggests that amphipod 

populations could persist at cadmium concentrations greater 

than the chronic criterion. An examination of data from 

the upper Clark Fork River near sites where Canfield and 

others (1994) collected amphipods showed that a maximum 

cadmium concentration of 1 μg/L reported in 1989 at a 

hardness of 172 mg/L was 1.3 times the chronic criterion. 

However, since 1993, when detection limits for the USGS 

dataset were lowered from 1 to 0.1 μg/L, the maximum 

cadmium concentration for the areas of the Clark Fork 

River where amphipods were collected was only 0.2 μg/L. 

The presence of amphipods in the Clark Fork River near a 

site where measured cadmium concentrations exceeded the 

chronic cadmium criteria once during the 3-year period prior 

to their collection is consistent with the concept that aquatic 

communities can tolerate infrequent exceedences of criteria. 

However, the evidence from this Clark Fork River case study 

is weak because the quality of routine USGS trace metals 

sampling and analyses prior to about 1993 was not as rigorous 

as that for subsequent data, and the more recent data show low 

cadmium concentrations. Hyalella were not abundant at any 

of the Clark Fork River sites sampled in 1991 (≤0.5 percent) 

(Canfield and others, 1994). In 1999, Hyalella were collected 

from tributaries to the Clark Fork River, but none were 

collected from several sites in the mainstem Clark Fork River 

(U.S. Geological Survey, 2005). Although failure to collect 

a taxa does not mean the taxa was absent from the sampled 

water body, it does indicate that the taxa was not common at 

the times and places sampled.

The Coeur d’Alene and Clark Fork River examples of 

amphipods being present in environments with cadmium 

concentrations near or exceeding the chronic criterion that 

was not necessarily designed to protect Hyalella suggests 

that in field conditions, Hyalella populations may be resilient 

to criterion concentrations. However, because the empirical 

data are limited, the potential effects of cadmium at chronic 

criterion concentrations to Hyalella populations were further 

evaluated through population modeling.

Simulating Effects of Cadmium to a Natural, Coldwater 

Hyalella azteca Population

With a GMCV of 0.33 μg/L versus the final chronic value 

0.41 μg/L cadmium at a hardness of 50 mg/L, Hyalella azteca

cannot be considered fully protected by the criteria. Based on 

the results of individual toxicity tests, continuous exposure 

to cadmium at the chronic criterion concentrations would 

result in adverse effects ranging from a 10 percent decrease in 

reproduction to greater than 50 percent mortality (table 11). 

However, in ecological risk assessments, the questions of 

interest usually relate to effects on the abundance, production, 

and persistence of populations and ecosystems, rather than 

the death or impairment of individual organisms. Effects of 

Table 11. Effects of cadmium to the most chronically sensitive species, Hyalella azteca, at chronic criterion concentrations.

[Abbreviations: CCC, criterion continuous concentration; ECp, percentage (p) of test population adversely affected at a concentration; LOEC, lowest observed 

effect concentration; CaCO3, calcium carbonate. Symbols: mg/L, milligram per liter; μg/L, microgram per liter; >, greater than]

Test No.1

(table 16)

Hardness 

(mg/L as 

CaCO3)

Chronic 

value 

(μg/L)

Chronic 

criterion 

(CCC, μg/L)

Percentage of 

organisms adversely 

affected at the CCC

Type of adverse effect Reference

45 17 0.16 0.20 40 Increased mortality Suedel and others, 1997

46 280 .98 1.0 10 Decreased number of 

young/females

Ingersoll and Kemble, 2001

47 126 .74 .65 25 Increased mortality Chadwick Ecological 

Consultants, Inc., 2004a

48 153 1.02 .73 20 Increased mortality Chadwick Ecological 

Consultants, Inc., 2004a

49 130 .25 .66 >50 Increased mortality Borgmann and others, 1991

50 130 .72 .66 10 Increased mortality Borgmann and others, 1989b

Mean 26 

1

Except for the tests numbered 45 and 49, ECp values that corresponded to criterion concentrations at test hardness were calculated from the authors’ data 

using a triangular probability distribution analysis for dichotomous mortality responses or by piecewise tailed regression analysis. ECp analyses were made 

using U.S. Environmental Protection Agency’s Toxicity Response Analysis Program (Erickson, 2002). For test 45, the ECp value is the effect at the reported 

LOEC, which at 0.25 μg/L was close to the criterion concentration of 0.20 μg/L calculated for the hardness of the test water. Insufficient data were reported 

to calculate a response curve. For test number 49, insufficient information was reported to estimate an ECp value that corresponded with the CCC. However, 

because the 42-day EC50 (0.53 μg/L) was less than the CCC, the CCC would have likely resulted in >50 percent mortality for this test.

50 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
contaminants at these “higher” levels of organization can be 

extrapolated through mathematical simulations that integrate 

individual-level toxicity testing results with ecological theory. 

Population-level effects of cadmium at its chronic criterion 

concentration to Hyalella azteca were evaluated using lifestage or age structured matrix projection models, also known 

as Leslie matrix models. The model uses a matrix of survival 

and reproduction rates (vital rates) of the different life stages, 

and projects them in time using matrix algebra. The technique 

basically is a bookkeeping device in which the births and 

deaths are tabulated over a given time frame (Barnthouse, 

1993; Caswell, 2001). 

Population Modeling Methods

The model was constructed in two stages. Stage 1 was to 

construct a population model that reasonably reflected baseline 

conditions, absent any toxicant effects. Stage 2 was to overlay 

effects of cadmium observed through lifecycle toxicity testing 

to the baseline survival and reproduction rates of Hyalella 

azteca. To build the baseline model, a life history model 

for was constructed for a Hyalella azteca population that 

would reflect temperate, coldwater conditions. Vital rates and 

demographic characteristics for the model were taken from 

field and laboratory studies of the natural history of the species 

(Cooper, 1965; Mathias, 1971; Strong, 1972; de March, 1978; 

Gibbons and Mackie, 1991). The modeled population was 

considered to have the following annual life cycle: eggs are 

deposited in late spring and early summer, followed by an egg 

and juvenile stage lasting about three months, followed by an 

adult stage that overwinters and then reproduces in late spring 

and early summer. During the reproductive stage, females 

mate and produce about four broods (one every 14 days for 

2 months) before dying. 

This annual life cycle seemed appropriate to assume 

for the mid-latitude vicinity of Idaho, but it would not be 

relevant throughout the species range. In North America, 

Hyalella azteca occur from at least Guatemala (latitude 16°N) 

to the Arctic (latitude 68°), as long as the mean summer 

water temperature of lakes reaches 10°C. In warmer regions 

of the species range, it has generation times on the order of 

6–8 weeks with shorter than annual life cycles (de March, 

1977; Pennak, 1978). However, assuming an annual life 

cycle simplified the population model and was consistent 

with observations from coldwater populations experiencing 

mean summer temperatures around 15–20°C (Mathias, 1971; 

Strong, 1972; de March, 1977, 1978). 

The demographic parameters and vital rates used in the 

modeling are summarized in table 12. Survival and fecundity 

in the model were further reduced at different cadmium 

concentrations by superimposing mortality and reduced 

reproduction from the results of Ingersoll’s and Kemble’s 

(2001) full life cycle study. This study was selected because of 

the seven chronic tests with Hyalella azteca that were located 

(table 16), it was the only one for which the concentration 

responses for the reproduction and survival endpoints were 

available. Because other studies have shown that reproduction 

in Hyalella azteca is strongly related to growth (Cooper, 1965; 

Strong, 1972; Ingersoll and others, 1998), results of population 

modeling using reproductive impairment also should have 

relevance to shorter “chronic” toxicity tests with Hyalella 

azteca that measured growth responses but were not continued 

through the full life cycle. 

Forecasts of the effects of cadmium on Hyalella azteca

populations used RAMAS® computer software (Spencer 

and Ferson, 1997a, 1997b). A transition matrix for the 

forecasts was built by recalculating the life stage based-vital 

rates as biweekly age-classes. Uncertainty in vital rates and 

environmental variation were estimated by running 500 

replicate simulations of each scenario, with each replicate 

having variable or uncertain parameters sampled from the 

most appropriate statistical distribution (that is, Monte Carlo 

simulations). 

Three basic scenarios were simulated. 

1. Population growth rates (lambda, λ) calculated 

for a range of constant cadmium concentrations. 

Population growth rates have been considered 

more ecologically relevant than single test 

endpoints because they integrate both survival and 

reproduction (Forbes and Calow, 1999). However, 

they only can be calculated for deterministic 

projection matrices where parameters are considered 

to be known exactly and are constant through time, 

which is unrealistic for real populations. 

2. Risk of population decline statistics were calculated 

across a range of constant cadmium concentrations. 

In contrast to the first deterministic scenario, these 

scenarios recognized the intrinsic unpredictability 

of natural systems (stochasticity) and used Monte 

Carlo simulations based on parameter variability 

estimates. Since these simulations produce a tangle 

of possible population trajectories, these projections 

are summarized with statistics describing the risks 

of the population declining by a given percentage at 

some time during the simulations. 

3. Seasonal population trajectories for several years 

under

a. Baseline conditions without any added cadmium 

stress; 

b. Adding constant exposure to cadmium at the 

chronic criterion; and 

c. An episodic exposure scenario in which the 

animals were exposed to 1.2 times the chronic 

criterion at the beginning of their reproductive 

cycle for 0.5 months once every 3 years with 

cadmium at 0.6 times the chronic criterion at 

other times. 

Results 51
This scenario is only slightly more severe than would be 

allowable under the frequency, magnitude, and duration for 

chronic criterion (the allowable duration is 4 days rather than 

1 month). The scenario also is similar to that observed in the 

Spokane River, Idaho, where average cadmium concentrations 

were about half that of the maximum concentrations and 

where maximum concentrations slightly exceeded the chronic 

criteria annually (fig. 3). This scenario was too complex to 

model directly, so the model was run in four phases with the 

outputs from one phase used as the initial conditions of the 

next phase.

Coincidentally, at the 280 mg/L hardness value of the 

toxicity test results used for the simulations, the chronic 

criterion is 1 μg/L so that multiples of criteria exceedence can 

be interpreted directly as concentrations, for example, 0.6X = 

0.6 μg/L. The models were projected for 6 years, that is, six 

generations with an annual life cycle. 

A lognormal distribution was assumed for variability 

in baseline fecundity rates, based on patterns observed in 22 

populations in lakes in Ontario (Gibbons and Mackie, 1991). 

Making realistic estimates of environmental variability was 

more problematic. Mathias’s (1971) mortality rates for a 

Hyalella population that was preyed on by salamanders and 

salmonids in a Northwestern lake seemed most relevant for the 

simulations; however, only mean mortality rates per station 

were reported. For juvenile Hyalella in a Michigan lake during 

the summer, Cooper (1965) reported weekly mortality rates 

ranging from -2 to 60 percent with an average of 26 percent 

Table 12. Vital rates used for modeling effects of cadmium to a coldwater Hyalella azteca population.

[Abbreviations: Cd, cadmium; EC50, 50 percent of test population. Symbols: mg/L, milligram per liter]

Parameter Mean Standard 

deviation Range Parameter Mean Standard deviation Range

Juvenile survival rates (overall 

for life stage)1

0.2825 0.25 – 0.32 Adult daily mortality rate during 

spring/summer (deaths per 

animal per day)1

0.0063 0.003 – 0.009

Juvenile half-monthly survival 

rates2

.908 0.22 0 – 1 Adult half-monthly survival, 

“A”2

0.9 0.05 0 – 1

Brood size, young per fecund 

female, “Y”3

8.35 2.28 6.3 – 15.5 Sex ratio, females:male, “S”3 1.7 : 1 .32 1.1 – 2.4

Proportion of females that are 

fecund, “F”3

.26 .07 0.14 – 0.430 Adult overwinter survival rate 

per half-monthly time step4

0.99

Broods per half-month time step 

during reproductive season 

“B”5

1 Fecundity (=A·Y·S·F·B)6 1.23 .3 0.44 – 4.26

Broods per life time 4 Ricker density dependence 

function for fecundity (β)

-0.00105 -0.00138 – 

-0.00071

Cd 42-day EC50 at 280 mg/L 

hardness7

2.49 Cd reproductive impairment 

(broodsize) EC50 at 280 mg/L 

hardness7

1.96

Slope of the probit curve at the 

EC507

.515 Slope of the probit curve at the 

EC50

7

0.313

1

Mean and range of observed survival at shallow and medium depth stations (Mathias, 1971).

2

Mean and range of survival rates adjusted from the duration of observations to the half-monthly time steps for the life stage used in the model (assuming 

3-months duration for juvenile life stage, 3 months for adult life stage, and 6-months duration for an inactive, overwintering phase).

3

Gibbons and Mackie, 1991.

4

Placeholder survival rate was assumed in order to adjust the model’s time steps to an annual life cycle. While true survival rates during winter are 

undoubtedly lower, overwinter survival is probably high, based on Cooper’s (1965) observations that mortality rates approached zero as fish predation declined 

with the onset of winter in a Michigan lake.

5

Assuming a 14-day resting period between broods, based on de March’s (1978) experimental results at 18oC and assuming a 2-month reproductive season 

during late spring and early summer.

6

Standard deviation for this compound parameter was estimated by calculating a coefficient of variation (CV) for each parameter and taking the product of the 

CVs.

7

EC50 values and the slope of the probit curve were estimated using EPA’s (1992) 2-parameter probit distribution using the raw data from the test described by 

Ingersoll and Kemble (2001).

52 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
and a coefficient of variation of 72 percent. For adults during 

summer, calculated mortality rates ranged from 43 to 552 

percent (Cooper, 1965). Although negative mortality rates 

can be assumed to be zero and reflect measurement error, 

how to interpret mortality rates greater than 100 percent was 

not obvious. Because real populations are influenced by 

unpredictable combinations of environmental factors such as 

weather, competition, and predation and sampling methods 

have considerable measurement error, ignoring variability 

and uncertainty in mortality rates would be environmentally 

unrealistic. Further, projections under static (deterministic) 

scenarios showed that the population structure was more 

sensitive to survival rates than to fecundity rates, so including 

some estimate of variability in mortality rates seemed 

appropriate. Thus, Cooper’s mean coefficients of variation 

for juvenile mortality rates were used to indicate the relative 

variability in mortality rates that would be plausible for a 

natural Hyalella population. These coefficients of variation 

were multiplied by Mathias’s mean values to obtain a plausible 

value for environmental variability in baseline mortality 

rates for a natural population. For adult mortality rates, the 

standard deviation of mean mortality rates from Mathias’s 

different stations was used for the Monte Carlo simulations, 

recognizing that this probably under represents the natural 

variability in adult mortality rates. 

Population Modeling Results

For the first scenario modeled under static, baseline 

conditions, the annual population growth rate (λ, lambda) 

for the population was calculated as between 1.01 and 1.09 

(fig. 4A). This indicates that the model population was near 

equilibrium and suggests that the input parameters were 

plausible. A population with a λ value of 1.0 is at equilibrium 

with births and deaths replacing each other. Absent any 

compensating factors, if λ is less than 1.0, the population 

will eventually decline to extinction. If λ is greater than 1.0, 

the population will eventually explode exponentially until 

resources become limiting. As cadmium concentrations 

increase above 100 percent of the chronic criterion, (1 μg/L 

under all modeling scenarios), λ values fall below 1.0. 

The second set of scenarios shows the likelihood of 

Hyalella population declines of at least 75, 90, 95, and 

98 percent (fig. 4B). This range of declines encompasses 

declines that would be common under baseline conditions (75 

percent) to declines that are severe enough to approximate a 

population extinction (98 percent). The choice of a 98 percent 

reduction as an quasi-extinction threshold was arbitrary, but 

probably reflects a decrease that is severe enough that even 

if the population did not die out completely, it would take 

a long time to rebound, absent immigration or extremely 

strong compensatory increases in survival and fecundity. 

The projections show that under baseline conditions, odds 

are even that a Hyalella population will experience at least 

75 percent declines at some point over 6 generations. More 

severe declines greater than 90 percent are unlikely (5 percent 

risk) with almost no risk of 95 percent or greater declines in 

abundance. The risk curves are nearly flat from 0–0.8 μg/L, 

increase slightly from 0.8–1.2 μg/L, and increase sharply at 

greater than 1.2 μg/L cadmium. At 1 μg/L (1 times CCC) the 

risk of a 90-percent decline has increased to 12 percent and the 

risk of 95 percent or greater decrease still is 1 percent or less. 

Above 1 μg/L, the risk curves for severe population declines or 

extinction increase steeply, so that at sustained concentrations 

of 1.9 μg/L, severe declines are nearly certain and extinction 

(estimated as a >98 percent decline) is probable. At 1.9 μg/L, 

decreases in abundance of at least 95 percent are forecast with 

a 96-percent probability. 

This 1.9 μg/L concentration is significant because it 

corresponds with several endpoints from the underlying 

toxicity test used in the simulations (Ingersoll and Kemble, 

2001). In the 42-day toxicity test, the EC50 for reproduction 

and both the LOEC and EC20 values for mortality were about 

1.9 μg/L, and the NOEC after 10-days exposure also was 

1.9 μg/L. Using the raw data from Ingersoll’s and Kemble’s 

(2001) test, the EC20 for growth at 28 days was estimated 

through logistic regression at about 1.7 μg/L and for 42 days 

at >3.2 μg/L. NOEC, LOEC, and EC20 statistics from chronic 

test endpoints have been considered acceptably low levels 

of toxic effects by some authors (see section “Statistical 

Interpretation of Chronic Tests”). The results of the population 

forecasts indicate that if such “low-effect” concentrations were 

persistent, disproportionably severe declines or even local 

extinctions of Hyalella populations could potentially occur.

The third set of scenarios forecasted seasonal trends 

in Hyalella abundances through six generations (fig. 4C). 

Under the baseline scenario (natural population with no 

added cadmium stress), the modeling predicted generally 

increasing populations with seasonal abundance cycles of 

as much as 3 times. Over the six-generations forecast, the 

troughs (annual minima) in the abundance cycles increased by 

about 35 percent per generation. Under a scenario of episodic 

exposures to 1.2 times the CCC every third generation at the 

beginning of the reproductive season, the population also 

trended upward but only by about 15 percent per year. Under 

a scenario of constant exposure to 1 times the CCC, little or 

no growth in the population was predicted. Of the scenarios 

modeled, the episodic exposure scenario probably is the one 

most relevant to potential effects of cadmium to Hyalella

under the magnitude, frequency, and duration of allowable 

criteria exceedences. The timing of the assumed exceedences 

mattered. Assuming that the exceedences occurred at the 

beginning of the reproductive season reduced reproductive 

potential more than did exceedences late in the reproductive 

season, or than exceedences during non-reproductive times of 

the year (not shown).

The seasonally cyclical abundance patterns produced 

by the model are qualitatively similar to patterns observed in 

seasonal field studies, although the patterns in the field studies 

were not as regular as in the model. (Cooper, 1965; Mathias, 

1971). The model likely underestimates interannual variability. 

Hyalella densities measured over a 12-year period at a USGS 

Results 53
id91U05_fig04

% Hyalella

Hyalella/m2

Snake River at King Hill, ID

USGS 13154500

Precentage of decline

75

90

95

Quasi-extinction

 Greater than 98%

Baseline

0.6X to 1.2X CCC episodic exposures

1X CCC continuous exposure

0

0.2

0.4

0.6

0.8

1.0

1.2

0 0.5 1 1.5 2 2.5

0

1

10

100

1,000

10,000

0

1

1

2

2

3

3

4

1992 1994 1996 1998 2000 2002 2004 2006

0

500

1,000

1,500

2,000

2,500

3,000

0

10

20

30

40

50

60

70

80

90

100

0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0

0 1 2 3 4 5 6

Cadmium, in micrograms per liter

Probablility of population decline, in percent

Population growth rate (lambda, 

λ)

Cadmium, in micrograms per liter

Females per square meter

Time in years Year sampled

Hyalella per square meter

Percentage hyalella

A. B.

C D. .

Figure 4. Risks to Hyalella azteca populations exposed to the cadmium chronic criterion. A. Population growth rate 

(lambda, λ) decline with increasing cadmium concentrations. Solid lines show λ for using min. and max. survival and 

fecundity estimates. Horizontal and vertical dashed lines show threshold for population decline (λ<1) and chronic criterion 

respectively; B. Risks of population declines up to the specified percentages occurring at some point during the 6-year 

population projections (mean probabilities following 500 replicate simulations); C. Projected Hyalella azteca densities 

(mean of 500 replicates) under scenarios of baseline conditions; exposure to 0.6X the chronic criterion concentrations 

(CCC) with episodic increases to 1.2X CCC at the beginning of the reproductive season for one 15-day time steps, once 

every 3 years ; and continuous exposure to 1.0X the chronic criterion concentration; D. Measured Hyalella azteca densities 

over a 12-year period or record at a USGS NAWQA biological trends monitoring station.

54 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
NAWQA long-term biomonitoring site in the Snake River, 

Idaho, ranged from 0 to 4707 per m2 (fig. 4D). Although some 

of this apparent variability undoubtedly reflects measurement 

error, these USGS NAWQA data probably were close to 

the best case in regard to minimizing uncertainty owing to 

differences in sampling and processing. All samples shown in 

figure 4D were collected using the same methods, equipment, 

and all samples were analyzed by the same laboratory, which 

for the most part used the same taxonomists during this 

time. Most of the samples also were collected by the same 

individuals and from the same riffle year-to-year. 

The population modeling forecasts of the effects of 

chronic cadmium exposure to Hyalella azteca populations 

could under- or over-predict the effects for several reasons. 

First, the toxic effects of cadmium are not constant across 

populations or conditions. The hardness-adjusted mortality 

observed by Ingersoll and Kemble (2001) was lower than 

mortalities observed in some other long-term exposures of 

Hyalella to cadmium (table 11). Although the tests resulting 

in higher mortalities were not used directly in the simulations 

because reproductive endpoints were not reported, this 

suggests that some populations could be more vulnerable to 

cadmium than those used in the present simulations. 

Second, the modeling assumed that the growth of 

Hyalella populations were density-independent. Densitydependence refers to the fact that populations cannot 

increase without limits, because expanding populations are 

all eventually self-limiting. Population growth is limited by 

density-dependent factors such as food and space limitations, 

waste buildup, disease transmission, or large populations 

may attract more predators. In population modeling, some 

mathematical function is applied to either or both fecundity 

or survival so that as densities increase or decrease, survival 

and fecundity rates also increase or decrease in compensation. 

Mathematically, the density-dependent compensatory 

changes in vital rates tends to lessen the effects of toxicants 

in population forecasts. For long-term population forecasts, 

assuming density-independence, as was the case here, is 

unrealistic. However, for shorter term forecasts, reasonable 

projections and comparisons can be made while ignoring 

density dependence (Barnthouse, 1993; Caswell, 2001). 

Further, even among intensely studied populations, the ability 

to reconstruct from data the density-dependence relations 

governing the natural dynamics may be limited because of 

changing environmental conditions. Under moderate or weak 

density dependence, the computed extinction risks are lower 

when density dependence is included in the model. Therefore, 

when available data sets are insufficient for reconstructing 

reliable measurements of density dependence, conservative 

estimates of extinction probabilities can be made from 

models that simply omit density dependence (Ginzburg and 

others, 1990). In Hyalella, weak density-dependent reduction 

in broodsizes has been observed in laboratory cultures 

(Borgmann and others, 1989b), however Strong (1972) 

observed that over the ranges encountered in nature, gross 

reproductive rates of individual Hyalella were uninfluenced 

by density. Strong (1972) suggested that resource-limited 

density-dependence is expected only in environments free of 

vertebrate predators. These predators appear generally able 

to crop Hyalella populations back to levels where resources 

do not limit reproduction. Predation that was biased against 

larger amphipods was implicated as the selective force shaping 

the small size, slow individual growth, and hence reduced 

fecundities of some populations (Strong, 1972). 

Some of the field studies were consistent with the idea 

that density-dependent compensation probably occurs with 

Hyalella populations. Cooper (1965) observed that Hyalella

populations in a Michigan lake could tolerate as much as 

about 28 percent mortality per week. This mortality rate is 

considerably higher than the mortality rates used here (about 

5 percent mortality per week) that resulted in a slowly growing 

baseline population (fig. 4C). Strong (1972) observed periodic 

mass mortality of Hyalella in a hot springs population that 

were apparently due to intrusions of noxious subsurface 

water. The periodic mass mortalities were frequent enough 

that the hot springs population spent most of its time below 

resource-determined carrying capacity (Strong, 1972). This 

recovery from acutely toxic episodes by an actual population 

complements the episodic chronic exposures to a model 

population (fig. 4C). Finally, the dramatic increase in Hyalella

density in Snake River samples from 1999 to 2003 of 0 to 

3,766 animals/m2 (fig. 4D) indicates that Hyalella populations 

can grow quickly under favorable environmental conditions. 

High flows in the Snake River during the late 1990s may 

have reduced and displaced Hyalella, and in following years 

the flows were low and stable and accompanied by lush 

macrophyte beds. 

A regression of mean Hyalella brood sizes and 

abundances that were reported by Gibbons and Mackie 

(1991) did result in a significant negative relation, suggesting 

density-dependent compensation in fecundity was present. 

This relation was used with the Ricker recruitment function 

to consider the effects of density dependent compensation in 

the population modeling. The Ricker function can be used 

to model situations where the sharing of resources among 

individuals is roughly equal. Therefore, as the population 

density becomes very high, no individual gets enough 

resources, and at high densities, all individuals suffer equally. 

This is sometimes described as “scramble competition” 

(Ricker, 1975; Spencer and Ferson, 1997b; Caswell, 2001). 

The Ricker function is P = exp(-βA) where P is the multiplier 

applied to maximum fecundity, A is the abundance in females 

per square meter, and β is a measure of the strength of density 

dependence in square meters per female. For fecundity, β

Results 55
was estimated by regression as -0.001046 with 95-percent 

confidence intervals of -0.00138 to -0.00071). Extinction risk 

forecasts were sensitive to the density-dependence function. 

In figure 4B, the risk curves for severe population decline or 

quasi-extinction begin to bend upward at around 0.9 μg/L 

and by about 2.0 μg/L, severe declines are almost inevitable 

and quasi-extinction is probable. If density-dependent 

compensation in fecundity is incorporated into the model as 

described above, the risk curves only begin to bend upward at 

around 1.6 μg/L. By 2.0 μg/L, instead of an 80-percent risk of 

a 98-percent population decline, there was about an 80-percent 

risk of a 75-percent decline and very little risk (<1 percent) of 

severe declines of greater than 90 percent (density-dependent 

results are not graphed).

However, the physiological mechanisms responsible for 

reproductive failure from cadmium exposure and reduced 

reproduction associated with high densities in the absence 

of toxic effects are different. Reduced reproduction in 

natural environments at high densities is likely related to 

factors such as reduced growth from competition for food or 

space, or from selective predation on large reproductively fit 

individuals. These factors would not seem relevant to reduced 

reproduction occurring in toxicity tests that are conducted 

in environments with ample food and no predation. No 

theoretical or experimental basis was obvious for why densitydependent compensatory increases in fecundity should be 

expected to offset cadmium-caused reproductive impairment 

equally. Because giving density-dependence and toxicity the 

same mathematical weight in adjusting the vital rates in the 

population simulations did not seem warranted, and because 

no basis was apparent for some other weighting, only the 

results of the density-independent population simulations are 

presented in figure 4. 

However, the lack of experimental testing of density 

dependence of Hyalella in combination with exposures to 

toxicants and the inability to describe mathematically densitydependent compensation in polluted environments does mean 

that none exists. Such compensation seems most plausible 

in episodic exposure scenarios. It follows that under these 

scenarios, the actual effects of cadmium exposures should be 

less severe than the density independent simulations carried 

out here.

In summary, the empirical field data and population 

modeling suggest that for Hyalella azteca, a species that is 

both the most sensitive species tested to date with chronic 

cadmium exposures as well as a species that by design is 

not necessarily protected by the chronic cadmium criteria, 

populations could be limited by cadmium at the chronic 

criterion. These population limitations are unlikely to be of a 

magnitude that would result in local extinctions. In montane 

stream food webs, Hyalella probably are not dominant 

enough for invertivores such as salmonids, sculpin, dace, or 

salamanders to depend on them as a food source. However, 

in springbrooks (and presumably other streams) that have 

stable, low velocity flows, Hyalella can dominate benthic 

macroinvertebrate communities. A crash of a prey species 

in such a system could have large effects on predatory fish 

populations. Hyalella and other amphipods could be important 

components of lake food chains, and a contaminant-caused 

population crash could have important adverse ecological 

effects. 

Field Validation

The goals of the Clean Water Act include maintaining 

and restoring the biological, chemical, and physical integrity 

of the waters of the United States. Because the goal of 

chemical integrity is central to the reasons for developing 

aquatic life criteria, it seems logical to evaluate the 

acceptability of chemical criteria by whether water allowed to 

have chemicals at criteria conditions would compromise the 

“biological integrity” of waters. Frey (1977) defined waters 

that have biological integrity as those that have 

 “... a species composition, diversity, and functional 

organization comparable to that of natural habitats 

of the region.” 

Frey’s definition has been widely used in the biological 

condition of waters including development of indexes of 

biological integrity (IBIs). IBIs are empirical additive indexes 

consisting of various metrics describing an assemblage such 

as species composition, diversity, functional organization, 

presence or absence of pollution sensitive or tolerant taxa 

(Karr, 1991). This definition is used in this field validation 

section to evaluate whether apparent effects that co-occurred 

with cadmium concentrations at or less than criteria 

concentrations were “unacceptable.”

EPA’s National guidelines were developed on the 

premise that effects that occur on a species in appropriate 

laboratory tests generally will occur on the same species in 

comparable field situations (Stephan and others, 1985). When 

enough species test values are available to describe a species 

sensitivity distribution, the distribution is at least implicitly 

considered to represent the distribution of sensitivities in 

natural ecosystems. The species sensitivity distribution 

of species mean values can be visualized as a cumulative 

distribution function (fig. 5). Assuming that the chronic 

distribution curve approximates responses in field settings, 

at sustained cadmium concentrations of less than 0.4 μg/L, 

less than 5 percent of species are predicted to be adversely 

affected; at about 3 μg/L, 50 percent are predicted to be 

adversely affected, and above about 25 μg/L, 90 percent of all 

species are predicted to be adversely affected. Comparisons 

to field studies can help evaluate if these predicted species 

responses seem to represent “unacceptable” or indiscernible 

effects.

56 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
id91UO5_fig05

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

1.0

0.1 1 10 100 1,000 10,000 100,000

Fraction of species affected

Species mean acute values (SMAV), in micrograms per liter

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

0.1 1 10 100

Fraction of species affected

Species mean chronic values (SMCV), in micrograms per liter

Final acute value

Cutthroat

trout

Rainbow trout

Final chronic value

Hyalella azteca

Daphnia magna

Mottled sculpin

Rainbow trout

Fountain darter Brook trout

Ceriodaphnia dubia

Brown trout

Aplexa hypnorum

Daphnia pulex

Flagfish

Fathead minnow

Aelosoma

Daphnia magna

Hyalella azteca

Fathead minnow

Ceriodaphnia dubia

Paper pondshell

Daphnia pulex

Mayfly Baetis tricaudatus

Figure 5. Species sensitivity distributions of cadmium values showing species mean values (open circles), ranges 

of species values (horizontal lines), 5th percentile final acute and chronic values (vertical dashed lines), and their fitted 

logistic distribution curves. All values adjusted to a hardness of 50 milligrams per liter CaCO3.

Results 57
An approach to determine if criteria would protect natural 

aquatic communities is to compare the occurrence of apparent 

instream effects to criteria exceedences at that location. The 

absence of apparent effects at sites that do not often exceed 

criteria, or the presence of apparent effects at those sites where 

criteria are frequently exceeded would support the relevance of 

the criteria. The converse of either would call in question the 

protectiveness or relevance of criteria.

Limitations, however, are in the ability to “validate” 

criteria in this manner. “Validation” is the process of 

comparing the overall result or output of a method, toxicity 

test, or model with observed effects in natural systems (Cairns 

and others, 1995). In this case, the criteria that resulted from 

many toxicity tests, calculations and decisions are compared to 

field observations of ambient biological conditions and criteria 

exceedences. Toxicity test results can never truly be validated 

or refuted based on field comparisons and vice versa. Toxicity 

tests are conducted in ecologically unrealistic environments, 

where all variables other than that being tested are held 

constant. By this means, causality can be assigned to the test 

variable. In field conditions, multiple biological, physical, and 

chemical variables interact. With many variables changing, 

strict causality may be difficult to establish.

The best possible outcome when seeking to “validate” 

toxicity test results with field observations is that similar 

types of apparent effects are observed in the field as resulted 

in toxicity tests at similar concentrations (Chapman, 1995; 

Clements and Kiffney, 1996; de Vlaming and Norberg-King, 

1999; Clements and others, 2002; Suter and others, 2002a). 

For example, if concentrations of cadmium caused mortality 

to cutthroat trout relative to controls in toxicity tests, and if 

similar concentrations were measured in a stream location 

with low abundance of cutthroat trout relative to abundance 

at reference conditions with low cadmium concentrations 

but otherwise similar habitat conditions, then the toxicitytest predictions would be considered “validated.” Although 

powerful, there are limitations to these comparisons. First, 

the absence of instream effects can never be proven; effects 

may be present that are too subtle to detect by field surveys. 

Second, proving that the presence of apparent effects was 

caused by the stressor of interest may be difficult, because 

other unmeasured or correlated variables could be the cause. 

Thus, here biological effects in lakes and streams are qualified 

as “apparent” effects. 

Cadmium Whole-Lake Ecosystem Experiment

In the late 1980s, Canadian researchers began a 

remarkable series of whole-lake ecosystem studies with 

cadmium at the Experimental Lakes Area of western Ontario. 

The purpose of the studies was to gain knowledge on the 

behavior and effects of metals in real aquatic ecosystems. 

At the time, most information on the effects of cadmium on 

aquatic organisms was based on single-species laboratory tests 

and field observations. In contrast, Canadian water-quality 

guidelines were intended for ecosystems, with a national 

goal of no observable effects on aquatic life over the long 

term with particular attention to the protection of sensitive 

species and life stages. To investigate whether the chronic 

water-quality guidelines that were based on single-species 

tests were protective in the wild, a 6-year study of the fate 

and effects of experimental cadmium enrichment on Lake 382 

of the Experimental Lakes Area was undertaken (Schindler, 

1988; Lawrence and others, 1996; Malley, 1996). The wholelake studies are particularly relevant for the present analysis 

because (1) the extrapolation of results of single-species 

laboratory tests to the wild is a question; (2) the concentrations 

of cadmium actually achieved in Lake 382 were close to the 

present hardness-adjusted chronic cadmium criteria values 

(0.153 to 0.185 μg/L, which were about 1.25 to 1.5 times the 

chronic criterion 0.12 μg/L calculated for a hardness of 

8 mg/L); and (3) prior to the cadmium additions, Lake 382 

was in a near-pristine condition so that changes in aquatic 

populations and communities from baseline or reference-lake 

conditions could be attributed to the experimental cadmium 

additions. The last point is in contrast to field surveys of 

polluted systems because these tend to have co-occurring 

mixtures of metals and other disturbances, greatly reducing the 

ability to assign causes to effects. 

 Background cadmium concentrations in Lake 382 

before cadmium additions were about 0.0016 μg/L (1.6 ng/L). 

Cadmium was added to the epilimnion during the ice-free 

seasons such that concentrations were gradually ramped 

up for the first 3 years of additions, and were maintained 

through the ice-free season at 0.153 to 0.185 μg/L for 3 years. 

When the lake was ice covered, no additions were made and 

concentrations decreased, so annual average concentrations 

were about 33–60 percent of the concentrations maintained 

during the ice-free season. Recovery was monitored for 

an additional 3 years after the cadmium additions ended 

(Lawrence and others, 1996). 

Previous ecosystem manipulations with various stressors 

indicated that among the earliest of responses to stress are 

changes in species composition of small, rapidly-reproducing 

species with wide dispersal powers such as phytoplankton, 

and the disappearance of sensitive organisms from aquatic 

communities (Schindler, 1987). However, during or following 

the near-criterion cadmium additions, no overt effects were 

detected on phytoplankton, zooplankton, macrobenthos, 

or fish assemblages. Population monitoring of crayfish 

(Orconectes virilis), lake trout (Salvelinus namaycush), white 

sucker (Catostomus commersoni), and minnows detected no 

differences from baseline or reference conditions attributable 

to cadmium treatments. 

58 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Bioaccumulation occurred at every trophic level. 

Concentrations stabilized in short-lived Hyalella at 17X 

background but in both the long-lived floater mussel 

Pyganodon grandis (whole body) and lake trout (kidney) 

concentrations reached 5–9X and 4–5X, respectively, after 6 

years but were still trending upward when cadmium additions 

were discontinued. Except for metallothionein induction, 

biomarkers of stress (for example, biochemical, histological, 

physiological, and behavioral) monitored in the large fish and 

floater mussels were unremarkable. Malley (1996) speculated 

that if the experiment had run longer, adverse effects likely 

would have later become evident, based on the increasing 

trends in fish tissue residues and sediment contamination. 

The monitoring of Hyalella azteca and lake trout 

populations in Lake 382 are of particular interest because 

adverse effects were observed with these species in laboratory 

exposures at concentrations less than the chronic criterion 

derived in this report. Although the inability to detect effects 

to Hyalella populations does not demonstrate that no-effects 

occurred, it does indicate that no population crash of Hyalella

occurred. With lake trout, Scherer and others’ (1997) “black 

box” experiment found that in a tank with cadmium exposed 

lake trout and rainbow trout fingerlings, more rainbow trout 

remained at the end of the test than in the control tank. From 

this they inferred that cadmium exposures reduced the ability 

of lake trout to capture prey (Scherer and others, 1997). The 

foreseeable implications of an impaired ability to capture prey 

include reduced growth or condition factor, which in turn 

implies less energy available for reproduction and lower fat 

reserves for winter, resulting in expected population declines 

and a shift in age class structure. Although none of these overt 

effects were obvious in cadmium-dosed Lake 382, this pattern 

of effect was clearly documented through similar monitoring 

efforts in acidification and other studies at the Experimental 

Lakes Area by the same team of researchers (Schindler and 

others, 1985; Schindler, 1987, 1988). Therefore, it appears 

unlikely that the lake trout exposed to elevated cadmium 

exposures in the wild for 6+ years in Lake 382 had appreciably 

impaired ability to capture food. These differences between 

tank experiments and ecosystem experiments are parallel to 

those observed with zooplankton. In limnocorral mesocosm 

exposures in preparation for the whole-lake experiments, 

marked reductions in the cladocerans Daphnia galeata and 

Holopedium gibberum were observed following exposure for 

3 weeks to 0.2 μg/L cadmium, yet these predictions of adverse 

effects in resident cladocerans in Lake 382 were not borne out 

(Lawrence and Holoka, 1991; Malley, 1996). 

Co-Occurrence of Cadmium Criterion Exceedences and 

Apparent Effects in Streams

The vicinity of the Coeur d’Alene River, Idaho has 

recently been the focus of several environmental assessments, 

and large chemical and biological datasets have been collected 

in the study area in the last several years. Here, chemical and 

biological data over a gradient of conditions were compiled 

and matched to seek associations between criteria exceedences 

and apparent adverse effects. The primary dataset evaluated 

was from the USGS National Water Quality Assessment, 

Northern Rockies and Intermontane Valleys study unit 

“NROK data” (http://id.water.usgs.gov/nrok/index.html). 

Two main strengths of this dataset for comparing instream 

conditions to criteria exceedences were that (1) chemical, 

physical habitat, macroinvertebrate, and electrofishing 

collection methods were synoptic, clearly described, suitably 

rigorous, and consistent; and (2) sample sites were selected 

to include a range of conditions from nearly undisturbed 

reference sites to highly disturbed mining sites. Additional 

biological and chemical data from the IDEQ and fish surveys 

conducted by natural resource trustees as part of a natural 

resource damage assessment were also used. These data were 

originally compiled in Dillon and Mebane (2002) and were 

presented in more detail there.

Several biological metrics (that is biological endpoints 

that are expected to respond to a stressor in a predictable 

way) were compared to cadmium concentrations. Because 

an almost limitless number of biological endpoints could 

potentially be examined, the comparisons focused on six 

metrics that were expected to be sensitive to metals, and could 

be calculated with the available data: trout density, percent 

sculpins, a stream fish index of biological integrity (IBI), total 

invertebrate taxa richness, mayfly taxa richness, and density of 

metals intolerant taxa. Brief descriptions of each metric follow 

and the apparent responses associated with chronic cadmium 

criteria exceedence factors are shown in figure 6. Biological 

metrics values are plotted with cadmium chronic criterion unit 

exceedences where an exceedence factor

EF Cd

Cd CCC = g/L

 g/L

( )

( )

μ

μ

was calculated for the hardness the sample. If no sample 

hardness were available, the CCC was calculated with the 

average hardness for the location. The interpretations of 

these comparisons were limited to (1) if there was no adverse 

response apparent for a metric at exceedence factors less than 

1.0, then sub-criterion cadmium concentrations did not cause 

an “unacceptable” adverse effect to that biological metric in 

this dataset; and (2) if there was an apparent adverse response 

threshold for a metric at an exceedence factor, then cadmium 

exceedences greater than that threshold could cause an 

“unacceptable” adverse effect to that biological metric in this 

dataset. However, in the latter case, since zinc also occurred 

at these sites in roughly equi-toxic proportions to cadmium, 

it would not be valid to assign causality for effects solely to 

cadmium since the effects could have been caused by zinc, 

their combination, or other factors.

Results 59
Figure 6. Comparison of measured cadmium chronic criterion exceedence factors, and corresponding instream 

biological condition metrics. Points indicate exceedence factors at the time of biological sampling, or for sites not 

sampled concurrently, median exceedence factors. Error bars show the 5th and 95th percentile exceedence factors for 

sites with multiple values. Exceedence factors calculated after adjusting the criterion for the sample or median hardness 

values for each site. Data from streams in the Coeur d’Alene and St. Regis River basins, Idaho and Montana (Dillon and 

Mebane, 2002).

id91U05_fig06

0

10

20

30

40

Trout density, in number

per 100 square meters

Stream fish, indexes of

biological integrity

Cadmium exceedance factors Invertebrate taxa richness

0

20

40

60

80

100 Percent sculpin

0

20

40

60

80

100

0.01 0.1 1 10 100

0

2

4

6

8

10

Number of mayfly taxa

0

500

1000

1500

2000

2500

3000

3500

0.01 0.1 1 10 100

0.01 0.1 1 10 100 0.01 0.1 1 10 100 Density of metals intolerant taxa, in number per square meter

0.01 0.1 1 10 100 0.01 0.1 1 10 100

0

10

20

30

40

60 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Trout density—Salmonids were the most acutely sensitive 

organisms to cadmium in the available data, and were fairly 

sensitive in chronic exposures (lower one-half of the species 

sensitivity distribution, table 7; fig. 5). Thus, in locations with 

cadmium concentrations greater than those toxic in controlled 

testing, trout densities would be expected to be depressed. 

The chronic criteria are intended to delineate levels safe for 

sensitive organisms based on testing at the individual level 

of organization. In nature, the collective effects of toxicity 

to individuals should be reflected at the population level of 

organization. However, quantitative relations between trout 

populations and environmental quality are notoriously elusive. 

Relations between environmental variables and trout densities 

are often unclear because of interactions among water quality, 

habitat, species interactions, and management manipulations 

(Fausch and others, 1988; Rose, 2000; Mebane and others, 

2003).

In the Coeur d’Alene dataset, trout densities were 

consistently low at locations with cadmium exceedence factors 

greater than about 10 (fig. 6). At sites with lower exceedence 

factors, densities were variable. At sites with exceedence 

factors less than 1.0 and lower trout densities than some other 

sites, the variability in densities was presumably owing to 

factors other than cadmium such as physical habitats. Whether 

an apparent peak in the trout densities at sites with cadmium 

exceedence factors of about 2 to 5 is real or an artifact of the 

dataset is unclear. 

Percent sculpin—Sculpin were among the most sensitive 

organisms in both acute and chronic testing with cadmium 

(table 7). Sculpin are ubiquitous and often numerically 

dominant in mid-sized forest streams in the Pacific Northwest. 

Although they tend to become less dominant as streams 

increase in size to become rivers, they are almost always 

present and are usually common in unpolluted, coldwater 

rivers in the inland Northwest (Maret and MacCoy, 2002; 

Mebane, 2002b; Mebane and others, 2003). Sculpins and 

other small-bodied fish may be superior to large-bodied 

fish as “sentinel” species for detecting effects from discrete 

disturbances in open receiving water environments. Smaller 

fish species, such as cottids, exhibit limited mobility relative 

to many larger species and typically possess a smaller home 

range. Many small species also show territorial behavior, 

particularly in lotic systems. This characteristic increases the 

probability that a sentinel species will not move extensively, 

and the observed response of that species will more likely 

reflect the local environment in which it was caught. In 

addition, small fish species tend to be more numerous than 

larger, more predatory species, which facilitates sampling; 

they have a shorter life span and therefore show alterations in 

reproduction and growth faster than longer-lived species; and 

they are not subject to commercial or sport fishing (Gibbons 

and others, 1998; Munkittrick and McMaster, 2000).

Sculpin showed a remarkably consistent threshold 

response to elevated cadmium (fig. 6). Sculpin were abundant 

at almost all sites where the cadmium criterion seldom 

exceeded 2 times. Sculpin were almost completely extirpated 

at sites where the median cadmium concentrations exceeded 

about 2 times cadmium chronic criterion. Because cadmium 

and zinc occurred in this study area in roughly similar 

proportions (on a toxic unit basis) at many sites the extirpation 

of sculpin could have been caused by either cadmium or 

zinc. An examination of the raw data behind figure 6 showed 

that among the sites at which sculpin were absent and at 

which median zinc exceedence factors were <1.0, the lowest 

cadmium exceedence factor was 2.4.

Index of biological integrity —The index of biotic 

integrity (IBI) concept was developed to address the need 

for operational definitions of Clean Water Act terms such as 

“biological integrity” and “unreasonable degradation.” The 

IBI was intended to provide a broadly based and ecologically 

sound tool to evaluate biological conditions in streams, 

incorporating many attributes of stream communities to 

evaluate human effects on a stream and its watershed. Those 

attributes cover the range of ecological levels from the 

individuals through population, community and ecosystem 

(Karr, 1991). An assumption behind a multimetric additive 

index such as the IBI is that at least some metrics would 

respond to most stressors. The IBI framework was used to 

develop an index for coldwater forest streams in Idaho that 

gauges a stream against an expectation based on minimal 

disturbance in the ecoregion. The IBI developed for fish 

communities in Idaho streams is an additive index consisting 

of the following six metrics: (1) coldwater native species; (2) 

percent coldwater individuals; (3) percent sensitive native 

individuals; (4) trout age classes; (5) sculpin age classes, and 

(6) catch per unit effort of coldwater individuals (Mebane, 

2002b).

Most IBI scores at sites with median cadmium 

exceedence factors less than 1 were high, indicating the 

fish community composition at study sites that usually met 

cadmium criteria was similar to that at reference streams. 

IBI scores showed a general graded decrease with further 

increases in cadmium exceedence factors greater than 2.0.

Macroinvertebrate taxa richness—Total taxa richness 

measures are widely used in field assessments of streams and 

may be sensitive measures of ecosystem disturbance. Many 

studies have reported declines in taxa richness in response 

to elevated metals concentrations in streams (Clements and 

Kiffney, 1996; Clements and others, 2000; Fore, 2002; Kiffney 

and Clements, 2002; Mebane, 2002a; Maret and others, 2003). 

Carlisle and Clements (1999) found that in terms of sensitivity, 

variability, and statistical power, richness measures were 

superior to other commonly used metrics. 

Results 61
In the Coeur d’Alene vicinity data, a general pattern of 

declining invertebrate richness corresponding to increasing 

exceedences of cadmium criteria is apparent (fig. 6). The 

sites with the highest species richness all had exceedence 

factors of 1.0 or less, the sample with the highest richness (38 

taxa) was collected from a site with an exceedence factor of 

about 1.0, and sites with exceedence factors of about 2.0 or 

more tended to have lower species richness. This suggests a 

possible exceedence factor response threshold between 1.0 

and 2.0. Overall, the invertebrate richness and exceedence 

factor comparison indicates that at locations where the chronic 

cadmium criterion is not exceeded, few if any taxa would be 

lost owing to cadmium toxicity.

Mayfly taxa richness—This richness measure is limited to 

mayfly (Ephemeroptera) taxa. As a group, mayflies generally 

are sensitive to metals and fewer taxa tend to be found in areas 

with elevated metals, and some studies have reported that 

the effects of metals generally were greater on mayflies than 

other macroinvertebrate groups (Clements and Kiffney, 1996; 

Clements and others, 2000; Fore, 2002; Kiffney and Clements, 

2002; Mebane, 2002a; Maret and others, 2003). Similarly to 

the overall taxa richness metric, a general pattern of declining 

mayfly richness corresponding to increasing exceedences of 

cadmium criteria is apparent. The highest number of mayfly 

taxa (9), we collected from a site with an exceedence factor of 

about 3 (fig. 6).

Metals-intolerant macroinvertebrate density—Several 

taxa were consistently sensitive to heavy metals in surveys 

of Colorado mountain streams (Clements and others, 2000; 

Fore, 2003). This metric measures the abundance of only 

these metals-intolerant taxa. Abundance is measured instead 

of taxa richness in this metric, because taxa richness reflects 

only presence of a taxon. The metric consists of the sum 

of the densities of the following 10 genera in 4 orders: the 

mayflies Cinygmula, Drunella, Epeorus, Paraleptophlebia, 

and Rhithrogena; the stoneflies Skwala, Suwallia, and Sweltsa; 

the caddisfly Rhyacophila, and the dipteran Pericoma. 

Conceptually, this metric should be more sensitive than taxa 

richness metrics because if a taxon is severely depressed but 

still present in reduced numbers, taxa richness counts will 

not reflect that. However, for the data analyzed in figure 6, 

this metric appears less sensitive than the richness measures. 

Densities of the 10 “metals-intolerant” taxa were >1,500 per 

square meter at sites with exceedence factors of about 8 or less 

(fig. 6).

In summary, for these streams which included a 

gradient of cadmium concentrations, at sites where the 

cadmium chronic criteria values derived in this report were 

seldom exceeded (EF <1.0) the macroinvertebrate and 

fish assemblages seemed to have “species composition, 

diversity, and functional organization comparable to that of 

natural habitats of the region.” In other words, for these data 

biological integrity was unlikely to have been compromised 

owing to cadmium concentrations so long as the chronic 

criterion derived in this report was seldom exceeded.

A second case study of the co-occurrence of cadmium 

criterion exceedences and apparent effects in streams gives 

some insight to the effects of a single pulse of cadmium to a 

stream macroinvertebrate community. In April 1999, a pulse 

of cadmium of about 30X background concentrations and with 

a chronic criterion EF of about 2.6 and an acute criterion EF 

of about 1.0 was detected in Thompson Creek, Idaho (fig. 3). 

Because the episode was detected through weekly effluent 

monitoring, the duration of the pulse is assumed to have 

been longer than 1 day and shorter than 1 week. By August 

1999 when the benthic macroinvertebrate community was 

sampled, few if any adverse effects were apparent. Overall 

macroinvertebrate density and taxa richness were similar at 

the exposed and upstream site. Densities and taxa richness of 

mayflies were somewhat higher in the downstream (exposed) 

sampling site than the upstream reference site, a pattern 

which also was observed at two companion sites located at 

similar elevations on a nearby stream that was not influenced 

by the release. The mayfly Heptagenia was only collected 

upstream of the exposed area in Thompson but was collected 

from both sites on the nearby unaffected stream (Chadwick 

Ecological Consultants, Inc., 2000). Because Heptagenia have 

been reported to be sensitive to elevated metals in streams 

(Clements and others, 2000; Fore, 2002; Clements, 2004), it is 

plausible that their absence from the survey was related to the 

earlier pulse. However, the distributions of macroinvertebrates 

in streams may be both spatially and temporally patchy, 

and the possibility that the correspondence between the 

distribution of Heptagenia and the cadmium pulse was merely 

coincidental cannot be excluded. Further, other mayflies in 

the family Heptageniidae that have also been considered 

sensitive to metals showed no such pattern. The Heptageniid 

mayflies Cinygmula and Epeorus were abundant upstream 

and downstream. No corresponding fish community data were 

collected in the August 1999 survey. Overall, this episode 

suggests that the macroinvertebrate communities in coldwater, 

mountain streams similar to Thompson Creek may be resilient 

to or recover quickly from, infrequent cadmium pulses with 

durations of few days and a magnitude as much as about 

2.5 times the chronic criterion concentration and 1 times the 

acute criterion concentration.

Other comparisons of chronic criterion values and 

apparent effects values from ecosystem studies and field 

surveys are given in table 13. Some comparisons were 

ambiguous, but most of the comparisons indicated the onset 

of adverse effects occurred at cadmium concentrations greater 

than the relevant chronic criterion concentrations. Two studies 

reported clearly adverse effects at sub-criterion concentrations. 

Juvenile hatchery brook trout exposed to cadmium for 30 days 

at 50 percent of the chronic criterion had reduced success at 

capturing mayfly prey. Riddell and others (2005a) concluded 

that impaired foraging by affected individuals has the potential 

to alter the structure of aquatic communities by modifying 

energy flow and trophic interactions. When combined with 

increased activity levels, as observed in exposed brook trout, 

the energy budget of such organisms could become further 

62 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
compromised, with potentially negative effects on growth, 

survival, and reproduction (Riddell and others, 2005a, 2005b). 

Whether brook trout or other salmonids exposed to subcriterion cadmium concentrations in field conditions also have 

impaired foraging is ambiguous. Brook trout were reasonably 

common in the co-occurrence surveys from the Coeur d’Alene 

vicinity described previously. No obvious reductions in several 

indictors of growth, survival, and reproduction were observed 

in trout surveys from sites at which cadmium concentrations 

were only moderately elevated (chronic exceedence factors 

less than about 2) (Dillon and Mebane, 2002; Maret and 

MacCoy, 2002). Lake trout populations that were exposed 

in the wild to higher cadmium chronic criterion exceedence 

factors than were the brook trout tested by Riddell and others 

(2005a, 2005b) showed no obvious decrease in condition or 

numbers (Malley, 1996; Vince Palace, Freshwater Research 

Institute, Winnipeg, oral commun., 2005). The absence of 

discernable effects of cadmium with lake trout in a long-term 

whole lake study and the presence of adverse effects in shorter 

term tests with brook trout could be related to the exposure 

and testing differences or species differences. Brook and lake 

trout are closely related, although their life histories differ 

markedly. Tests with lake trout in simplified food chains have 

shown adverse effects at similar cadmium concentrations 

causing the adverse effects in the brook trout studies 

(Behavioral Toxicity section; table 17). This suggests differing 

cadmium exposures or test differences during whole-lake or 

field studies may be more important than species differences 

for resolving the apparently contradictory results with brook 

trout and lake trout. 

In summary, the “field validation” comparisons 

showed mostly good agreement between the laboratorybased predictions and effects observed in the field surveys 

or ecosystem experiments (fig. 7). Adverse effects at 

concentrations lower than the calculated chronic criterion 

were reported from small scale, simplified model ecosystem 

experiments. In these experiments, adverse behavioral effects 

were observed with two species in the genus Salvelinus, brook 

trout and lake trout. In a large-scale ecosystem study and field 

surveys, no obvious adverse effects to lake trout or brook trout 

were observed at cadmium chronic criterion concentrations. 

The disparity between effects observed in small-scale or 

large-scale ecosystem experiments or surveys could be real or 

artifacts of differing study designs and data types.

Figure 7. Comparison of cadmium effects concentrations estimated from ecosystem studies or field assessments 

to proportions of adversely affected species based on a species-sensitivity distribution (SSD) of laboratoryderived species mean chronic values (SMCVs). Solid curved line illustrates a SSD fitted from SMCVs using logistic 

regression; dashed line indicates chronic criterion concentration (CCC), numbers in parentheses refer to studies 

described in table 13. All values were adjusted to a hardness of 50 milligrams per liter CaCO3.

Subtle benthic

community shifts

in an effluent

dominated urban

stream (9)

Loss of about 50% of

species in a stream

invertebrate assemblages (1)

Daphnia

magna

population

crashed (4)

Pronounced and

consistent adverse

effects to a cold

water fish

assemblages

(loss of about 50%

diversity), (1)

No adverse effects

detected to

phytoplankton,

zooplankton,

macrobenthos, or

fish assemblages

following wholelake Cd additions. (3) No adverse effects

detected in field surveys

of coldwater fish or

invertebrate stream

assemblages (1)

Adverse effects to

brook trout in an

experimental stream

food chain (8)

CCC, 5th percentile of the SSD

Increased mortality

to small mayflies in

an experimental

Colorado stream (10)

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

0.1 1 10 100

Fraction of species affected

Species mean chronic values (SMCV), in micrograms per liter

Results 63
Table 13. Comparison of chronic criterion values with results of ecosystem studies and field surveys.

[Underlined values are lower than corresponding chronic criterion values. Hardness as mg/L CaCO3, all other concentrations are for measured values in 

microgram per liter. Abbreviations: Cd, cadmium; Cu, copper; Pb, lead; L, liter; m, meter; NOEC, no observed effect concentration; LOEC, lowest observed 

effect concentration; CCC, criterion continuous concentration. Symbols: μg/L, microgram per liter]

Study type Study 

location

Exposure 

duration

Method 

notes Effects/Notes

Hardness 

(mg/L 

as 

CaCO3)

Lowest 

adverse 

effect 

value

Corresponding 

chronic 

criterion

 value

Reference

Field

assessments

Northern Idaho

and Western

Montana

Indefinite Co-occurrence 

comparisons

Extirpation of sculpin (Cottus sp.) from 

all sites where median cadmium 

concentration exceeded ≈ 2.4 times 

chronic criterion. Value listed was lowest 

concentration with an associated adverse 

effect (sculpin extirpated), which did not 

also have zinc criterion exceedences

64 1.20 0.44 Dillon and 

Mebane, 2002

Field

assessments

Western 

Kentucky

Indefinite Co-occurrence 

comparisons

Reduced macroinvertebrate taxa richness 

and density, particularly for mayflies. 

Cd body burden highest in the least 

resistant taxa (Stenonema mayflies). 

LOEC range shown is the mean and 

estimated maximum Cd concentrations 

(mean plus S.D. of 0.55 μg/L). Mean 

Cu concentrations ~0.5 times CCC and 

mean Pb concentrations ~1 times CCC, 

suggesting that the combined exposures 

may have influenced the observed effects

63 .32 to .87 .44 Birge and others, 

2000

Field

assessments

Northwestern 

Québec, 

Canada

Indefinite Co-occurrence 

comparisons 

sedimentwater 

interface

Floater mussel (Pyganodon grandis) 

population status was negatively 

correlated with modeled free-cadmium 

ion concentrations at the water-sediment 

interface. Cd concentrations in water were 

estimated from sediment core extracts 

using a geochemical model. Effects 

of Cd were confounded by variable 

temperatures, other natural factors, and 

marked decreases in metal contamination 

over the 10 years prior to sampling

15 .07 .19 Perceval and 

others, 2004

Lake 

experiment

Experimental 

Lakes Area, 

W. Ontario, 

Canada

6 years Epilimniondosed 

during icefree season

No overt effects detected on phytoplankton, 

zooplankton, macrobenthos, or fish 

assemblages. Bioaccumulation occurred 

at every trophic level. Concentrations 

stabilized in short-lived Hyalella at 17 

times background but in both the longlived floater mussel Pyganodon grandis

(whole body) and lake trout (kidney) 

concentrations both reached 9 times after 

6-years but were still trending upward. 

Increasing trends of fish bioaccumulation 

and sediment contamination suggesting 

effects might eventually occur in 

experiment had run longer. Cd value is the 

mean of the concentrations maintained 

during the final two ice-free seasons

8 .19 .12 Lawrence and 

others, 1996; 

Malley, 1996

64 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Study type Study 

location

Exposure 

duration

Method 

notes Effects/Notes

Hardness 

(mg/L 

as 

CaCO3)

Lowest 

adverse 

effect 

value

Corresponding 

chronic 

criterion

 value

Reference

Lentic 

mesocosm 

Burlington, 

Ontario (Lake 

Ontario water)

250 days 4.5 m deep × 1 m 

wide columns

Daphnia population crashed at 9 weeks, 

apparently could adapt to shorter 

exposures. No effect at 1 μg/L

130 4.70 0.69 Borgmann and 

others, 1989a

Lentic 

mesocosm

Experimental 

Lakes Area, 

W. Ontario

14 days In situ 

continuous 

flow vessels

Decrease in abundance of cladocerans 

Daphnia galeata and Holopedium 

gibberum

8 .20 .12 Lawrence and 

Holoka, 1991

Pond Blacksburg, 

Virginia

28 days Suspended 

colonization 

substrates

Decrease in protozoan assemblage 

colonization (species richness). NOEC 

0.4 and LOEC 

1.4 μg/L

70 .75 .47 Niederlehner 

and others, 

1985

Stream 

mesocosm

Saskatoon, 

Saskatchewan, 

Canada

30 days Recirculating 

stream 

channels

Juvenile brook trout success at capturing 

mayfly prey impaired. Activity by stonefly 

Kogotus nonus nymphs decreased. Mayfly 

Baetis tricaudatus drift increased

156 .4 .77 Riddell and 

others, 2005a

Stream 

mesocosm

Saskatoon, 

Saskatchewan, 

Canada

30 days Recirculating 

stream 

channels

Reduced condition factor in juvenile brook 

trout. Brook trout prey preference shifted 

from mayflies to less motile midges

156 .4 .77 Riddell and 

others, 2005b

Stream 

mesocosm

Denton, Texas 10 days Once-through, 

effluent 

dominated 

stream 

channels

Subtle benthic community shifts (increase 

in Chironomid abundance, decrease in 

aquatic moth larvae)

146 15 .74 Brooks and 

others, 2004

Stream 

mesocosm

Fort Collins, 

Colorado

10 days Once-through 

stream 

channels

Increased mortality to smaller individuals of 

the mayflies Baetis tricaudatus (Baetidae), 

Ephemerella infrequens (Ephemerellidae), 

and Rhithrogena hageni (Heptageniidae), 

and the stonefly Pteronarcella badia

(Pteronarcyidae)

29 3.20 .27 Kiffney and 

Clements, 

1996

Stream 

mesocosm

Aiken, South 

Carolina

91.5-m-long 

stream

Periphyton and invertebrate community 

structure affected. No effects to freeranging fish observed, reported value is 

the NOEC

9.1 .5 .13 Versteeg and 

others, 1999

Stream 

microcosm

Glen Lyn, 

Virginia

20-L volume, 

1.5-m-long 

stream

No observed effects on benthic microbial and 

invertebrate communities

57 2.5 .41 Versteeg and 

others, 1999

Table 13. Comparison of chronic criterion values with results of ecosystem studies and field surveys.—Continued

Results 65
Cadmium and Threatened and Endangered 

Species

The National guidelines for developing criteria include 

assumptions that species can tolerate some adverse effects, 

and that populations and ecosystems have resiliency and 

can probably recover from periodic adverse effects (Stephan 

and others, 1985, p. 8–12). These assumptions may not hold 

in cases of especially vulnerable species that are at risk of 

extinction through most of their range and that have less 

capacity to withstand or recover from additional stressors. 

These species include those listed as threatened or endangered 

under the Endangered Species Act (ESA). The goals of the 

ESA include conserving the ecosystems that endangered 

species and threatened species depend, to conserve those 

endangered and threatened species, and promoting their 

recovery. It is necessary that aquatic life criteria for cadmium 

or any substance be sufficiently protective that the criteria 

would not allow conditions that impede the recovery of 

threatened or endangered species. 

Risks of direct effects of cadmium on threatened or 

endangered species were evaluated using a simple risk 

quotient approach (table 14). In this approach, the highest 

no-observed-effect-concentration (NOEC) available for a 

species is the denominator, and the corresponding chronic 

criterion is the numerator. If the quotient is less than one, 

the criterion is presumed to be adequately protective, and if 

the quotient is greater than one, the criterion is presumably 

insufficiently protective (although the biological effect 

of exceeding a no-observed-effect value is admittedly 

ambiguous). 

Chronic test results were available for 3 of the 11 

threatened or endangered aquatic species that occur in Idaho. 

For most species for which no direct information is available, 

estimates need to be based on one or more taxonomic or 

toxicological surrogate species. For sockeye salmon and the 

Snake River physa, surrogate species were considered to 

be closely enough related that estimated cadmium NOECs 

were taken directly from the surrogate species, without any 

further adjustments. For white sturgeon, a two-step estimate 

of a cadmium NOEC was made. The relative sensitivity of 

three other sturgeon species and rainbow trout to copper, “a 

representative metal,” were compared (Dwyer and others, 

2005b). The most sensitive sturgeon response was 0.8 times 

as sensitive as rainbow trout to copper. The mean rainbow 

trout cadmium NOEC was then multiplied by this rainbowsturgeon sensitivity factor to obtain a proportional estimate of 

white sturgeon sensitivity to chronic cadmium exposures. In 

comparative testing of 17 diverse threatened or endangered 

species and standard toxicity test fish species, rainbow trout 

were consistently the most sensitive of the standard test fish 

species, and were the best estimate of sensitivity of cold- or 

warm-water threatened or endangered species (Sappington and 

others, 2001; Dwyer and others, 2005b). The rainbow trout 

was equal to or more sensitive than listed and related species 

81 percent of the time. Of the 19 percent of test results with 

listed species and rainbow trout in which the listed species 

were more sensitive, the geometric mean of the sensitivity 

factor between the rainbow trout and listed species was 0.63 

(Dwyer and others, 2005b). More conservative factors than the 

geometric mean can be determined using variance estimates 

(0.46 based on 1 standard deviation (SD) of the mean and 

0.33 based on 2 standard deviations of the mean). Using the 

most conservative of the sensitivity factors and assuming 

that the rainbow trout NOECs and LC50s are approximately 

proportional, in table 14 the rainbow trout NOEC is 

multiplied by 0.33 to estimate the low range of potential 

NOECs for snails (except for those in the families Physidae 

and Hydrobiidae, discussed separately). These potential 

NOEC values are shown as “greater than” values because the 

sensitivity factor of 0.33 is based on an extreme estimate of 

sensitivity differences, rather than a central tendency estimate. 

In a normally distributed population, the mean minus 2 

standard deviations corresponds to about the 2.5th percentile. 

Thus about 97.5 percent of the values would be greater than 

a value based on the mean minus 2 standard deviations. 

Using the most conservative factor as described, the 

approximate NOEC would be about equal to the CCC (0.37 

versus 0.38 μg/L, respectively). Using less conservative 

sensitivity factors of 0.46 based on 1 standard deviation of the 

mean, or 0.63 based on the geometric mean of the differences 

in the sensitivity factor between the rainbow trout and listed 

species, would result in NOEC estimates that are greater than 

the CCC (0.5 or 0.7 μg/L, respectively).

For Hydrobiidae snails, NOEC estimates also were 

made using interspecies correlation between fathead minnow 

(surrogate species) and Amnicola (predicted species) using 

EPA’s ICE model (Asfaw and others, 2004) with the fathead 

minnow mean chronic NOEC. Amnicola was a factor 14 more 

acutely resistant than fathead minnow to 5 chemicals. Fathead 

species mean NOEC was 4.5 μg/L at hardness of 50 mg/L. 

Assuming proportional acute and chronic responses, 

4.5 μg/L ́14 = 63 μg/L.

This latter estimate for Hydrobiidae snails probably is 

exorbitantly high, based on species-sensitivity distributions 

(fig. 4) and on chronic values reported with Physid or 

Lymnaeaid snails. The estimate is included simply to explore 

some various approaches to estimating chronic values for 

untested species. From the information reviewed here, the best 

estimate of a chronic NOEC value for all Snake River snails is 

probably around 0.5 to 0.7 μg/L at a hardness of 50 mg/L. 

66 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 14. Comparison of estimated no-effect concentrations for threatened and endangered species occurring in Idaho to chronic 

criteria concentrations.

[Abbreviations: NOEC, no observed effect concentration; CCC, criterion continuous concentration; LC50, 50 percent of life cycle; –, not applicable. Symbols:

mg/L, milligram per liter; μg/L, microgram per liter; <, less than; >, greater than]

Species of interest Family Hardness 

(mg/L)

Test 

NOEC 

(μg/L)

CCC 

Hazard 

quotient 

(CCC/NOEC)

NOEC 

basis

Family of 

surrogate Surrogate species

Bull trout, Salvelinus 

confluentus

Salmonidae 29 0.38 0.27 0.71 Direct – –

Chinook salmon, 

Oncorhynchus 

tshawytscha

Salmonidae 25 .96 .25 .26 Direct – –

Steelhead, Oncorhynchus 

mykiss1

Salmonidae 50 1.2 .38 .32 Direct – –

Sockeye salmon, 

Oncorhynchus nerka

Salmonidae 50 1.2 .38 .32 Surrogate Salmonidae Oncorhynchus mykiss

White sturgeon, Acipenser 

transmontanus2

Acipenseridae 50 .9 .38 .4 Surrogate Acipenseridae Atlantic sturgeon (Acipenser 

oxyrhynchus), Shortnose 

sturgeon (Acipenser 

brevirostrum), Shovelnose 

sturgeon (Scaphirhynchus 

platorynchus)

Bliss Rapids snail, 

Taylorconcha 

serpenticola3

Hydrobiidae 50 63 .38 .006 Surrogate Hydrobiidae Spire snail, Amnicola sp.

Idaho springsnail, 

Pyrgulopsis idahoensis3

Hydrobiidae 50 63 .38 .006 Surrogate Hydrobiidae Spire snail, Amnicola sp.

Bruneau Hot Springs snail, 

Pyrgulopsis bruneauensis3

Hydrobiidae 50 63 .38 .006 Surrogate Hydrobiidae Spire snail, Amnicola sp.

Desert valvata, Valvata 

utahensis4,5

Valvatidae 50 >.39 .38 <.97 Surrogate Salmonidae Oncorhynchus mykiss

Banbury Springs lanx, 

Lanx sp.5

Lymnaeidae 50 >.39 .38 <.97 Surrogate Salmonidae Oncorhynchus mykiss

Snake River physa, Physa 

natricina

Physidae 45 1.9 .35 .19 Surrogate Physidae Aplexa hypnorum

1

Geometric mean of six hardness-adjusted NOEC values.

2

In comparative testing, the lowest acute value of the three surrogate species tested to copper was 0.8 times the rainbow trout acute value to copper (Dwyer 

and others, 2005b). Despite their caution that complications with the sturgeon testing may have exaggerated their apparent sensitivity, they are used here in this 

comparison. Assuming acute-chronic responses are approximately proportional, the chronic rainbow trout NOEC was multiplied by 0.8 for the comparison here.

3

Estimate made using interspecies correlation between fathead minnow (surrogate species) and Amnicola (predicted species) using U.S. Environmental 

Protection Agency’s ICE model (Asfaw and others, 2004) with the fathead minnow mean chronic NOEC. Amnicola was 14 times more acutely resistant than 

fathead minnow to 5 chemicals. Fathead species mean NOEC = 4.5 μg/L at hardness 50. Assuming proportional acute and chronic responses, 4.5 μg/L times 

14 = 95 μg/L.

4

Also known as “Utah valvata.”

5

Rainbow trout used as a taxonomic-independent surrogate since in comparisons of 3 standard test species to 17 listed species for 5 chemicals representing a 

broad range of toxic modes of action, rainbow trout were more sensitive than listed species 81 percent of the tests. In tests with copper, a “representative” metal, 

rainbow trout were equal to or more sensitive than all the listed species tested (Dwyer and others, 2005b). Of the 19 percent of test results with listed species and 

rainbow trout in which the listed species were more sensitive, the geometric mean of the sensitivity factor between the rainbow trout and listed species was 0.63. 

Here, assuming that the rainbow trout NOECs and LC50s are approximately proportional, the rainbow trout NOEC is multiplied by 0.63 to estimate listed snail 

NOECs.

Results 67
Indirect Effects to Listed Salmonids Via Effects to 

Sensitive Prey Items

Indirect effects on listed species through food chain 

alterations are conceivable because by design, adverse effects 

could be allowed to as much as 5 percent of the taxa in an 

ecosystem. If a more sensitive taxa was disproportionately 

important in the food web of a listed species, indirect adverse 

effects could result to the listed species. The amphipod 

Hyalella azteca is the only taxa for which data were available 

to evaluate this potential scenario directly. Hyalella azteca’s

species mean chronic value was slightly lower than the 

calculated chronic criterion value (0.33 versus 0.41 μg/L 

cadmium), making it the most sensitive species tested to date 

with chronic cadmium exposures as well as a species that by 

design is not necessarily protected by the chronic cadmium 

criteria. 

Field data and population modeling suggested that 

Hyalella azteca populations could be limited by cadmium 

at the chronic criterion concentration. These population 

limitations are unlikely to be of a magnitude that would result 

in local extinctions. In montane stream food webs, Hyalella

probably are not dominant enough for invertivores such as 

salmonids, sculpin, dace, or salamanders to depend on them 

as a food source. However, in springbrooks and presumably 

other streams that have stable, low velocity flows, Hyalella

can dominate benthic macroinvertebrate communities. A crash 

of a prey species in such a system could have large effects 

on predatory fish populations. Hyalella and other amphipods 

could be important components of lake food chains, and a 

contaminant-caused population crash could have important 

adverse ecological effects (see section “Risks of the 5th 

Percentile Species-Sensivity Distribution (SSD)-Based 

Chronic Criterion to a More Sensitive Species: populationlevel effects to the amphipod Hyalella azteca”). 

These scenarios suggest that if cadmium concentrations 

in receiving waters approach chronic criterion values, 

biomonitoring could be prudent to evaluate if the diversity 

and abundance of potential salmonid prey items are similar to 

reference or baseline conditions. For example, a salmonid prey 

metric was defined for stream macroinvertebrates that could 

be useful for interpreting general biomonitoring when risks of 

indirect food chain effects to salmonids is a concern (Suttle 

and others, 2004).

Potential Site-Specific Adjustments to Cadmium 

Criteria Values

The information presented to this point suggested several 

scenarios where the criteria values derived could be under- or 

overprotective for a specific water body. Potential scenarios 

where the criteria values calculated in this report might be 

under- or overprotective, and possible steps to reduce the 

uncertainties in the degree of protectiveness include:

Possibly underprotective—In behavioral studies 

following extended exposures of salmonids in the genus 

Salvelinus to sub-criterion cadmium concentrations, reductions 

in the ability of salmonids to capture prey, shifts in prey 

utilization, and reductions in condition factors have been 

observed (Kislalioglu and others, 1996; Scherer and others, 

1997; Riddell and others, 2005b). These results suggest that 

if typical ambient cadmium concentrations were only slightly 

less than the chronic criterion concentrations, similar types of 

effects could occur in field conditions. However, reduced size 

or condition factors in Salvelinus were not obvious in two field 

situations with fish populations having long-term exposures 

to proportionally higher cadmium concentrations (Malley, 

1996; Maret and MacCoy, 2002). These apparent differences 

between laboratory and field responses could be real or they 

could simply reflect the difficulty in detecting biological 

changes in field studies that are small in relation to natural 

variability. If biomonitoring of fish communities in waters that 

receive cadmium discharges show shifts in prey utilization 

or decreases in body condition that are congruent with those 

observed by Riddell and others (2005b) in their experimental 

streams; it follows that the chronic cadmium criterion should 

be adjusted downward. The magnitude of such a potential 

adjustment is unclear, because Riddell and others (2005b) did 

not observe a threshold of effects.

Possibly underprotective—If a critical species at a 

site were sensitive to cadmium at lower than criterion 

concentrations, then a lower criterion it might be needed. For 

example, if cadmium were to be discharged to a waterbody 

in which the amphipod Hyalella azteca was an important 

component of fish diets, and in which Hyalella azteca

dominated the invertebrate assemblage such that alternative 

prey items were limited, then under these circumstances 

Hyalella azteca might be considered a critical species that 

warrants a downward adjustment to the chronic cadmium 

criterion. For example, a chronic criterion could be calculated 

based upon the species mean chronic value for Hyalella 

azteca, rather than the 5th percentile of the species-sensitivity 

distribution, as was done in this report. 

Possibly overprotective—Chemical characteristics of 

site water could make cadmium less toxic than in the typical 

test waters for the datasets that were used to derive criteria 

values in this reports. Most information reviewed indicated 

that cadmium toxicity is mitigated by increasing calcium 

concentrations. Since calcium concentrations are often 

correlated with other factors that may influence toxicity such 

as hardness, pH and alkalinity, to some extent these factors are 

often accounted for in the hardness-adjusted criteria values. 

However, these relations may not hold at all sites, and other 

factors that may influence toxicity such as high calcium to 

magnesium ratios or dissolved organic carbon concentrations 

may be important at some locations. Several tests have shown 

that cadmium is less toxic in effluent influenced stream waters 

than it is in reconstituted test waters (Diamond and others, 

1997; Brooks and others, 2004; Stanley and others, 2005). 

68 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
However this may not be a good comparison for site waters, 

because standard recipes for reconstituted water may produce 

poor surrogates for natural waters (Welsh and others, 2000; 

2001), and most of the data compiled in tables 15 and 16 of 

this report used waters originating from natural sources rather 

than reconstituted test waters.

Possibly overprotective—If the species occurring in 

a particular waterbody are likely less sensitive than those 

species that influenced the criteria derivation (that is, the four 

most sensitive genera), then the criteria values calculated 

here could be more stringent than needed to protect aquatic 

life. For example, the acute criterion value was lowered to 

protect critical species. If none of those species occur at a 

site that lowering may not be warranted. Similarly, two of the 

four most chronically sensitive genera were amphipods; if 

amphipods are unlikely to occur in a particular waterbody then 

the criteria might be more stringent than needed. 

U.S. Environmental Protection Agency (1994) suggests 

that for waters where sensitive species used in the general 

criterion calculation and comparably sensitive species do 

not occur, or a narrower mix of species than that found 

in the general criterion data occurs at a site, then it may 

be appropriate to recalculate the criterion after excluding 

nonresident species from the dataset. In Idaho and many other 

jurisdictions, protected aquatic life uses are often classified 

according to perceived thermal regimes, such as coldwater 

or warmwater fisheries. Accordingly, Chadwick Ecological 

Consultants, Inc. (2004b) classified organisms as “coldwater” 

or “warmwater” and recalculated cadmium criteria for 

warmwater uses after excluding coldwater taxa, and vice 

versa. In a related approach, in lieu of a thermal classification, 

Dillon and Mebane (2002) used a lotic/lentic classification 

of organisms for inclusion in site-specific criteria datasets 

for coldwater streams in northern Idaho. Because some taxa 

such as amphipods and cladocerans were eurythermal, but 

were only found in lentic (slow moving) waters, they were 

excluded from consideration in criteria datasets (Dillon and 

Mebane, 2002). Both approaches have some risk that by 

further limiting the diversity of the criteria dataset, sensitive 

but untested taxa might not be fully protected. Dwyer 

and others (2005b) compared the relative sensitivities of 

commonly tested fish species to 17 endangered and threatened 

species and concluded that rainbow trout data likely represent 

the response of sensitive warmwater species and not merely 

responses of coldwater species. Procedures that exclude 

species because of temperature would likely not be protective 

of sensitive warmwater species (Dwyer and others, 2005b). In 

a similar analysis, the standard effluent toxicity test species 

Ceriodaphnia dubia (a lentic species), was almost always 

more sensitive than either the endangered species or the 

fathead minnow in comparative testing with seven endangered 

fish species with diverse effluents. In contrast, the fathead 

minnow, a species that occurs in both lentic and lotic waters, 

was less sensitive than the endangered species in about 

80 percent of the comparisons (Dwyer and others, 2005a).

Summary

1. Cadmium is rare in aquatic environments, but it may be 

the most acutely toxic of all the atomically stable metals 

in the periodic table.

2. A large body of data on the effects of cadmium to aquatic 

animals has been reported. About 278, 93, and 102 

acute, chronic, and other, respectively, test values were 

summarized in this report. Because previous reviews 

indicated that aquatic plants tended to be less sensitive 

to cadmium than sensitive aquatic animals, data searches 

were focused on studies with aquatic animals.

3. Most of data on the effects of cadmium to aquatic 

species were generated from tests where cadmium was 

tested as the single contaminant of concern. However, 

in ambient waters, cadmium probably seldom occurs as 

a single contaminant of concern and commonly occurs 

with zinc and copper. Data on the toxicity of cadmium in 

chemical mixtures or interactive effects were equivocal. 

In tests with binary cadmium and zinc mixtures, most 

data indicate cadmium is no more toxic than it is when 

tested singly. Although toxicity data with cadmium and 

copper mixtures were contradictory, some data indicate 

that adding copper may result in greater toxicity than 

cadmium alone or than cadmium and zinc mixtures. 

4. Species mean acute values (SMAVs) and species mean 

chronic values (SMCVs) were determined for 69 and 28 

aquatic species respectively. These values were assumed 

to represent a subset or surrogates for the speciessensitivity distribution (SSD) of the approximately 1,200 

aquatic animal species in Idaho.

5. Data on the effects of cadmium through bioaccumulation 

into tissues residues were ambiguous. Tissue residue 

concentrations that co-occurred with the presence of 

observed adverse effects- and those that corresponded 

with the absence of observed adverse effects overlapped 

substantially. The data were insufficient to define 

maximum acceptable tissue concentrations and 

bioconcentration or bioaccumulation factors that could be 

used to back calculate water column concentrations that 

would reliably predict bioaccumulation risk.

6. Hardness-toxicity relations were determined that 

explained between 30–99 percent of the variability in 

toxicity for different species and data sets.

7. When normalized to a hardness of 50 mg/L as calcium 

carbonate, a Final Acute Value (FAV) and Final Chronic 

Value (FCV) were calculated from the 5th percentile of 

SSD models, based on genus mean values. The FAV 

is divided by 2.0 to extrapolate from a concentration 

that caused unacceptable adverse effects to acutely 

sensitive taxa (killed 50 percent) to a criterion maximum 

Summary 69
concentration expected to kill few if any organisms 

following acute exposures. In contrast, the FCV was 

calculated from chronic test values which reflected 

less severe effects than deaths of 50 percent of the test 

organisms. For fish, the chronic values corresponded with 

deaths or adverse effect to around 10–15 percent of the 

individuals, and for invertebrates around 20–30 percent. 

No further extrapolation of the FCV from low-effect to 

“no-effect” values was made.

8. Three species in the acute dataset had species mean acute 

values lower than the 5th percentile of the SSD (cutthroat 

trout, rainbow trout, and bull trout). The three species 

are the only species of trout native to Idaho and as the 

top carnivores in many coldwater ecosystems in Idaho, 

the three species were considered “critical” species. The 

species are also highly valued socially. Accordingly, the 

5th percentile of SSD-based FAV was lowered to protect 

these sensitive species.

9. One species in the smaller chronic dataset had a SMCV 

lower than the 5th percentile of the SSD (the amphipod 

Hyalella azteca). Unlike the case with native trout in 

the acute dataset, Hyalella azteca was not considered 

a priori to be a critical species. Instead, the potential 

population-level and food chain consequences of exposure 

to a chronic criterion that was not designed to protect 

Hyalella azteca were evaluated. A population model was 

constructed that simulated Hyalella azteca’s seasonal 

life cycle in temperate, coldwater ecosystems. A model 

of cadmium toxicity to Hyalella azteca was nested with 

the baseline population model to forecast the effects of 

continuous or episodic cadmium exposures. The forecasts 

suggested that extinctions of Hyalella azteca populations 

were unlikely, although population declines were 

expected.

10. The role of Hyalella azteca in aquatic food chains is 

probably proportional to its relative abundance in benthic 

communities. In a food chain study in an impounded 

section of the Snake River in Washington and Idaho, 

Hyalella azteca contributed about 3–8 percent of the diets 

of juvenile Chinook salmon and steelhead. If benthic 

communities are diverse and alternate prey species are 

abundant, declines in Hyalella azteca populations seem 

unlikely to have a disproportionate effect throughout 

aquatic food webs.

11. The chronic criterion concentration for cadmium that 

was calculated using the 5th percentile of a SSD based 

on single-species laboratory experiments was compared 

with the presence or absence of apparent effects due to 

cadmium from ecosystem experiments or field surveys. 

In general, there was reasonable agreement between 

the laboratory-based predictions and effects observed 

in the field surveys or ecosystem experiments. Adverse 

effects at concentrations lower than the calculated chronic 

criterion were reported from small scale, simplified 

model ecosystem experiments. In these experiments, 

adverse behavioral effects were observed with two 

species in the genus Salvelinus, brook trout and lake 

trout. In a large-scale ecosystem study and field surveys, 

no obvious adverse effects to lake trout, brook trout, or 

other fishes were observed at cadmium chronic criterion 

concentrations. The disparity between effects observed 

in small-scale or large-scale ecosystem experiments 

or surveys could be real or artifacts of differing study 

designs and data types.

12. In streams in and around Idaho, maximum weekly or 

maximum monthly cadmium concentrations tended 

to be about 2 times higher than long-term average 

concentrations, suggesting that if streams met the chronic 

criterion, long-term average cadmium concentrations 

would be around 50 percent of criteria concentrations 

at many sites. This suggests that in practice, the chronic 

criterion might be more protective than would appear 

solely based on species-sensitivity distributions, and 

assuming indefinite exposures. 

13. If ambient cadmium concentrations approached the SSDbased cadmium chronic criterion and if species that might 

not be fully protected are present, uncertainties in the 

criteria extrapolation and protectiveness could be reduced 

through monitoring of aquatic communities at exposed 

and reference sites. Field monitoring endpoints that 

correspond to experimental effects include the diversity 

and abundance of macroinvertebrate taxa, relative 

importance of prey items in the stomach contents of fish, 

condition factors of fish, fish species composition, age 

class distribution and relative abundances of fish species. 

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v. 78, p. 81–93.

Zhao, Y., and Newman, M.C., 2004, Shortcomings of 

the laboratory-derived median lethal concentration for 

predicting mortality in field populations: exposure duration 

and latent mortality: Environmental Toxicology and 

Chemistry, v. 23, no. 9, p. 2147–2153.

References Cited 87
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88 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.

[Method: D, “dissolved” test concentrations measured after being filtered through a 0.45-μm filter; F, flow-through test exposures; M, test concentrations 

measured during exposures; R, renewed test exposures; S, static test exposures; T, “total” metal concentrations determined from unfiltered samples; U, 

unmeasured test concentrations. LC50: Concentration killing 50 percent of test organisms. The term “LC50” as used here is equivalent to the term “EC50” 

as used by some authors to report results of acute tests where the only endpoints measured were mortality or immobilization, which was considered effective 

mortality. Italicized LC50 values may have been obtained with resistant life stages and were not used in SMAV calculations. SMAV: Species mean acute value.

GMAV: Genus mean acute value. Data different than EPA: “No” indicates that the test descriptions and values listed here for a test are identical to those listed 

in table 1a of Environmental Protection Agency (2001). “Yes” indicates that the values listed here either were not included in table 1a of U.S. Environmental 

Protection Agency (2001) or some aspect of the test descriptions or summary results in this table differs from those in table 1a of U.S. Environmental Protection 

Agency (2001). Abbreviations: EPA, U.S. Environmental Protection Agency; ACR, acute to chronic ratio; CaCO3, calcium carbonate; Cd, cadmium; DOC, 

dissolved organic carbon; LOEC, lowest-observed-effect concentration; MgSO4, magnesium sulfate; MN, Minnesota; MO, Missouri; mo, month; yr, year; 

oC, degrees Celsius; g, gram; hrs, hours; mg, milligram; mg/L, milligram per liter; mm, millimeter; ng, not given in source; μg/L, microgram per liter; μm, 

micrometer. Symbols: <, less than; ≤, less than or equal to; –, missing or inapplicable values; ~, approximately; ±, plus or minus]

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Actinonaias 

pectorosa

Pheasant shell 

(mussel)

Juvenile S,M,T 82 46 30.41 – – U.S. Environmental 

Protection Agency, 

2001, citing Keller, 

unpublished

1 No

Juvenile S,M,T 84 69 44.7 36.87 36.87 1 No

Aelosoma headleyi Polychaete 

worm

NonreproductiveS,M,T 62 1,200 1,002 – – Niederlehner and others, 

1984

– Yes

NonreproductiveS,M,T 168 4,980 1,806 1,345 1,345 – Yes

Ambystoma gracile Northwestern 

salamander

3-mo old larva F,M,T 45 468.4 511.57 511.57 511.57 Nebeker and others, 

1995

– No

Aplexa hypnorum Physid snail Adult F,M,T 45 93 101.01 101.01 101.01 Holcombe and others, 

1984

2 No

Adult F,M,T 45 93 101.01 – – Phipps and Holcombe, 

1985

2 Yes

Arctopsyche sp. Caddisfly Field collected R,M,D 30 467 716.06 716.06 716.06 Windward, 2002 – Yes

Asellus sp. Isopod ng F,M,T 220 2,129 616.26 616.3 616.26 U.S. Environmental 

Protection Agency, 2001

3 No

Baetis tricaudatus Mayfly Field collected R,M,T 156 1,611 621.74 – – Irving and others, 2003 4 Yes

Field collected R,M,D 21 73 150.86 306.3 306.26 Windward, 2002 5 Yes

Branchiura 

sowerbyi

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 240 1,569.61 1,570 1,569.61 Chapman and others, 

1982

– No

Carassius auratus Goldfish 8.8 g F,M,T 44 748 832.45 832.5 832.5 U.S. Environmental 

Protection Agency, 2001

– No

Catostomus 

commersoni

White sucker ng F,M,T 18 1,110 2,609.69 2,610 2,610 Environmental Protection 

Agency, 2001

– No

Ceriodaphnia 

dubia

Cladoceran, 

water flea

≤24hrs S,M,D 17 63.1 155.62 – – Suedel and others, 1997 6 Yes

<24 hrs R,M,T 100 27.3 15.29 – – Spehar and Fiandt, 1986 7 Yes

<24 hrs S,M,T 290 120 27.57 – – Schubauer-Berigan and 

others, 1993

8 Yes

<24 hrs R,M,T 80 54.5 36.78 – – Diamond and others, 

1997

– No

<24 hrs R,M,T 170 38.3 13.76 – – Brooks and others, 2004 – Yes

<24 hr S,M,T 40 31 37.94 32.86 32.86 Shaw and others, 2006 4 Yes

Chironomus 

riparius

Midge 2nd instar R,M,T 105 13,000 6,988 6,988 – Williams and others, 

1986

– Yes

Chironomus 

tentans

Midge 2nd instar, 

10–12 days 

S,M,T 17 8,000 19,730 19,730 11,742 Suedel and others, 1997 – Yes

Table 15 89
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Coregonus 

clupeaformis

Lake whitefish Yearlings, 

140 mm, 

22 g

F,M,T 81 530 354 354 354 McNicol, 1997 34 Yes

Cottus bairdi Mottled sculpin MO strain, 

newly 

hatched

F,M,T 101 2.9 1.62 – – Besser and others, 2007 – Yes

MO strain, 

swimup

F,M,T 101 5.2 2.91 – – – Yes

MN strain, 

swimup

F,M,T 101 3.6 2.02 – – – Yes

MN strain, 

0.03 g, 

3-week 

swimup

F,M,T 101 8 4.48 2.61 – 9 Yes

MN strain, 

0.1 g, 

7-week 

F,M,T 101 16.5 9.24 – – – Yes

MN strain, 

0.26 g, YOY

F,M,T 101 12.9 23 – – – Yes

MN strain, 

yearling

F,M,T 101 176 98.54 – – – Yes

Cottus confusus Shorthead 

sculpin

Field collected, 

30–60 mm 

TL, ~age 

1-2 yr.

R,M,D 21 1.29 2.67 2.67 2.67 Windward, 2002 – Yes

Cyprinella 

lutrensis

Red shiner ng S,M,T 86 6,620 4,225 4,225 4,225 U.S. Environmental 

Protection Agency, 2001

– No

Daphnia ambigua Cladoceran, 

water flea

<24 hrs S,M,T 40 10 12.2 12.2 – Shaw and others, 2006 4 Yes

90 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Daphnia magna Cladoceran, 

water flea

Genotype A S,M,T 170 3.6 1.29 – – Baird and others, 1991 – No

Genotype A-1 S,M,T 170 9.0 3.23 – – – No

Genotype A-2 S,M,T 170 9.0 3.23 – – – No

Genotype B S,M,T 170 4.5 1.62 – – – No

Genotype E S,M,T 170 27.1 9.73 – – – No

Genotype S-1 S,M,T 170 115.9 41.6 – – – No

<24 hrs, 

clone F

S,M,T 170 3.3 1.2 – – Barata and Baird, 2000 – Yes

<24 hrs, 

genotype A

S,M,T 170 3.6 1.29 – – Barata and others, 1998 – No

<24 hrs, 

genotype 

A-1

S,M,T 170 9 3.23 – – – No

<24 hrs, 

genotype 

A-2

S,M,T 170 9 3.23 – – – No

<24 hrs, 

genotype B

S,M,T 170 4.5 1.62 – – – No

<24 hrs, 

genotype E

S,M,T 170 27.1 9.73 – – – No

<24 hrs, 

genotype 

S-1

S,M,T 170 115.9 41.63 – – – No

<24 hrs, 

clone S-1

S,M,T 170 129.4 46.47 – – – No

Clone S-1 S,M,T 46 112 119.9 – – – No

Table 15 91
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Daphnia magna—

Continued

Cladoceran, 

water flea—

Continued

Clone S-1 S,M,T 91 106 64.40 – – Barata and others, 1998 – No

Clone F S,M,T 179 233 80.15 – – – No

<24 hrs S,M,T 179 23.6 8.12 – – – No

<24 hrs R,M,T 105 30 16.13 – – Canton and Sloof, 1982 – No

<24 hrs R,M,T 209 30 9.06 – – – No

<24 hrs S,M,T 51 9.9 9.74 – – Chapman and others, 

1980

h No

<24 hrs S,M,T 104 33 17.88 – – h No

<24 hrs S,M,T 105 34 18.28 – – h No

<24 hrs S,M,T 197 63 20 – – h No

<24 hrs S,M,T 209 49 14.81 – – h No

<24 hrs F,M,T 130 58 26.07 – – Attar and Maly, 1982 – No

<24 hrs S,M,T 170 37 13.29 – – Lewis and Weber, 1985 – Yes

<24 hrs S,M,T 170 6.1 2.19 – – – Yes

<24 hrs S,M,T 170 43 15.44 – – – Yes

<24 hrs S,M,T 170 31 11.13 – – – Yes

<24 hrs S,M,T 170 18 6.46 – – – Yes

<24 hrs S,M,T 170 12 4.31 – – – Yes

<24 hrs S,M,T 170 24 8.62 – – – Yes

<4 hrs S,M,T 76 59 41.56 – – Nebeker and others, 

1986a

– Yes

<4 hrs S,M,T 74 84 60.51 – – – Yes

<4 hrs S,M,T 41 99 116.88 – – – Yes

<4 hrs S,M,T 38 164 206.34 – – – Yes

<4 hrs S,M,T 76 71 50.01 – – – Yes

<4 hrs S,M,T 74 178 128.22 – – – Yes

<4-hrs S,M,T 74 116 83.56 – – – Yes

<4-hrs S,M,T 71 101 75.32 – – – Yes

1 day S,M,T 71 4 2.98 – – – Yes

1 day S,M,T 41 8 9.45 – – – Yes

1 day S,M,T 38 16 20.13 – – – Yes

1 day S,M,T 74 146 105.17 – – – Yes

2 days S,M,T 76 55 38.74 – – – Yes

2 days S,M,T 74 306 220.42 – – – Yes

2 days S,M,T 41 98 115.70 – – – Yes

2 days S,M,T 38 307 386.25 – – – Yes

2 days S,M,T 76 37 26.06 – – – Yes

2 days S,M,T 74 94 67.71 – – – Yes

2 days S,M,T 74 277 199.53 – – – Yes

2 days S,M,T 71 135 100.67 – – – Yes

92 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Daphnia magna—

Continued

Cladoceran, 

water flea—

Continued

5 days S,M,T 76 17 11.98 – – Nebeker and others, 

1986a

– Yes

5 days S,M,T 74 40 28.81 – – – Yes

5 days S,M,T 41 30 35.42 – – – Yes

5 days S,M,T 38 131 164.82 – – – Yes

5 days S,M,T 76 25 17.61 – – – Yes

5 days S,M,T 74 36 25.93 – – – Yes

5 days S,M,T 71 18 13.42 – – – Yes

5 days S,M,T 34 36 49.71 – – Nebeker and others, 

1986b

– Yes

5 days S,M,T 34 33 45.57 – – – Yes

5 days S,M,T 34 24 33.14 – – – Yes

5 days S,M,T 34 40 55.23 – – – Yes

<24 hrs S,M,T 40 101 121.9 – – Shaw and others, 2006 4 Yes

<24 hrs, 

clone F

S,M,T 170 24.5 8.80 – – Stuhlbacher and others, 

1992

– No

<24 hrs, 

clone S-1

S,M,T 170 129.4 46.47 – – – No

6 days, 

clone F

S,M,T 170 49.1 17.63 – – – No

<24 hrs old, 

clone F

S,M,T 170 25.4 9.12 – – Stuhlbacher and others, 

1993

– No

3 days, 

clone S-1

S,M,T 170 228.8 82.17 – – – No

6 days, 

clone S-1

S,M,T 170 250 89.79 – – – No

≤24 hrs S,M,T 17 26.4 65.11 – – Suedel and others, 1997 – Yes

≤24 hrs R,M,T 170 26 9.34 – – Ward and Robinson, 2005 – Yes

≤24 hrs R,M,T 170 34 12.21 – – – Yes

≤24 hrs R,M,T 170 39 14.01 – – – Yes

≤24 hrs R,M,T 170 48 17.24 – – – Yes

≤24 hrs R,M,T 170 55 19.75 – – – Yes

≤24 hrs R,M,T 170 63 22.63 – – – Yes

≤24 hrs R,M,T 170 100 35.92 – – – Yes

≤24 hrs R,M,T 170 120 43.10 22.29 – gt Yes

Daphnia pulex Cladoceran, 

water flea

<24 hrs S,M,T 53 70.1 66.76 – – Stackhouse and Benson, 

1988

– No

≤24 hrs R,M,T 58 115 101.57 – – Ingersoll and Winner, 

1982

10 Yes

≤24 hrs R,M,T 85 130 83.39 – – Lewis and Weber, 1985 37 Yes

≤24 hrs R,M,T 85 120 76.98 – – 37 Yes

≤24 hrs R,M,T 85 170 109.05 – – 37 Yes

≤24 hrs R,M,T 85 130 83.39 – – 37 Yes

Table 15 93
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Daphnia pulex—

Continued

Cladoceran, 

water flea—

Continued

≤24 hrs R,M,T 85 190 121.88 – – Lewis and Weber, 1985 37 Yes

≤24 hrs R,M,T 85 160 102.63 – – 37 Yes

≤24 hrs R,M,T 85 150 96.22 – – – Yes

≤24 hrs R,M,T 85 130 83.39 – – – Yes

≤24 hrs R,M,T 85 150 96.22 – – – Yes

≤24 hrs R,M,T 85 100 64.15 – – – Yes

≤24 hrs R,M,T 85 180 115.46 – – – Yes

≤24 hrs R,M,T 85 130 83.39 – – – Yes

≤24 hrs R,M,T 40 45 54.19 87.19 28.72 Shaw and others, 2006 – Yes

Dendrocoelum 

lacteum

Planarian ng R,M,T 87 24,702 15,540 15,540 15,540 U.S. Environmental 

Protection Agency, 2001

– No

Diporeia spp. Amphipod Juveniles, 

3.3 ± 1.3 mg

S,M,T 160 60 22.67 22.67 22.67 Gossiaux and others, 

1992

4, 36 Yes

Etheostoma 

fonticola

Fountain darter 4–6 days R,M,T 270 15.7 3.83 – – Castillo and Longley, 

2001

– Yes

4–6 days R,M,T 261 14.23 3.57 – – – Yes

4–6 days R,M,T 285 13.27 3.09 – – – Yes

4–6 days R,M,T 270 11.77 2.87 3.32 3.32 – Yes

Gambusia affinis Mosquitofish ng F,M,T 11 900 3,195.02 – – U.S. Environmental 

Protection Agency, 2001

– No

ng S,M,T 11 2,200 7,810.05 4,995.32 4,995.32 – No

Gammarus 

pseudolimnaeus

Amphipod ng S,M,T 44 68.30 76.74 76.74 76.74 Spehar and Carlson, 1984 – No

Gasterosteus 

aculeatus

Threespine 

stickleback

ng R,M,T 107 23,300 12,327 12327 12327 U.S. Environmental 

Protection Agency, 2001

– No

Glossiphonia 

complanata

Leech ng R,M,T 123 480 226.01 226.01 226.01 U.S. Environmental 

Protection Agency, 2001

12 No

Gyraulus sp. Snail larvae R,M,D 21 73 150.86 150.86 150.86 Windward, 2002 gt Yes

Hyalella azteca Amphipod Larger juveniles 

and young 

adults

S,M,T 34 8 11.05 – – Nebeker and others, 

1986b

– Yes

7–10 days S,M,T 48 3.8 3.93 – – Jackson and others, 2000 h, 32 Yes

7–10 days S,M,T 118 12.1 5.90 – – h, 32 Yes

2–3 weeks S,M,D 17 2.8 6.91 – – Suedel and others, 1997 h, 6, 13 Yes

0–2 days S,M,T 90 13 7.95 – – Collyard and others, 1994 h,14 Yes

2–4 days S,M,T 90 7.5 4.59 – – h,14 Yes

4–6 days S,M,T 90 9.5 5.81 – – h,14 Yes

10–12 days S,M,T 90 7 4.28 – – h,14 Yes

16–18 days S,M,T 90 11.5 7.03 – – h,14 Yes

24–26 days S,M,T 90 14 8.56 – – h,14 Yes

7–14 days S,M,T 290 5 1.15 5.01 5.01 Schubauer-Berigan and 

others, 1993

8 Yes

Hydra oligactis Hydra Monecious 

species

S,M,T 210 320 96.3 – – Karntanut and Pascoe, 

2002

15 Yes

94 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Hydra viridissima Hydra Monecious 

species

S,M,T 210 210 63.2 – – Karntanut and Pascoe, 

2002

15 Yes

Hydra vulgaris Hydra Zurich strain, 

male clone

S,M,T 204 310 95.6 – – Karntanut and Pascoe, 

2002

15 Yes

Zurich strain, 

male clone

S,M,T 210 520 156.5 – – 15 Yes

Dioecious 

strain

S,M,T 210 160 48.15 – – 15 Yes

Ictalurus punctatus Channel catfish 7.4 g F,M,T 44 4,480 4,985.77 4,985.77 4,985.77 Phipps and Holcombe, 

1985

– No

Jordanella floridae American 

flagfish

4-5 weeks F,M,T 44 2,500 2,782.24 2,782.24 2,782.24 Spehar, 1976 – No

Lampsilis 

straminea 

claibornensis

Southern 

fatmucket 

(mussel)

Juvenile S,M,T 40 38 45.80 45.80 45.80 U.S. Environmental 

Protection Agency, 2001

– No

Lampsilis teres Yellow 

sandshell 

(mussel)

Juvenile S,M,T 40 11 13.26 – – U.S. Environmental 

Protection Agency, 2001

– No

Juvenile S,M,T 40 33 39.77 22.96 22.96 – No

Lepomis cyanellus Green sunfish Juvenile S,M,T 86 11,520 7,353.46 – – U.S. Environmental 

Protection Agency, 2001

h No

Life stage 

unknown

S,M,T 335 20,500 4,173.80 5,540.03 5,540.03 h No

Lepomis 

macrochirus

Bluegill 40 g, 13 cm 

adults

S,M,T 18 2,300 5,407.47 – – Bishop and McIntosh, 

1981

h, 33 No

40 g, 13 cm 

adults

S,M,T 18 2,300 5,407.47 – – h, 33 No

Life stage 

unknown

F,M,T 207 21,100 6,426.94 – – U.S. Environmental 

Protection Agency, 2001

h No

1.0 g juvenile F,M,T 44 6,470 7,146.11 5,896.91 5,709.35 Phipps and Holcombe, 

1985

h No

Limnodrilus 

hoffmeisteri

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 170 1,111.81 – – Chapman and others, 

1982

h No

ng S,M,T 152 2,400 946.60 1,025.88 1,025.88 Environmental Protection 

Agency, 2001

h No

Lirceus alabamae Isopod ng F,M,T 152 150 59.16 59.16 59.16 – No

Lumbriculus 

variegatus

Aquatic 

earthworm

Mixed age 

adults

S,M,T 290 780 179.18 179.18 179.18 Schubauer-Berigan and 

others, 1993

– No

Lymnaea stagnalis Swamp 

lymnaea, 

pond snail

4-week 

juvenile

S,M,T 250 742 192.99 192.99 192.99 Coeurdassier and others, 

2004

– Yes

Oncorhynchus 

clarki lewisi

Westslope 

cutthroat 

trout

Swimup fry, 

Sandpoint 

strain

R,M,D 21 0.35 0.72 – – EVS Environment 

Consultants, 1996; 

Windward, 2002

16 Yes

Field collected 

YOY, 

20-50 mm 

TL

R,M,D 21 0.93 1.92 – – Windward, 2002 – Yes

Table 15 95
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Oncorhynchus 

clarki lewisi—

Continued

Westslope 

cutthroat 

trout—

Continued

Swimup fry, 

captive 

broodstock 

source, avg. 

wt.(g)= 0.2

R,M,D 32 1.41 2.05 – – Windward, 2002 – Yes

Swimup fry, 

captive 

broodstock 

source, avg. 

wt.(g)=0.18

R,M,D 31 1.18 1.78 1.50 – – Yes

Oncorhynchus 

mykiss

Rainbow trout Swimup 

(0.19 g, 

~3 weeks)

F,M,T 24 1.30 2.40 – – Chapman, 1978b h No

Parr (7 g, 

5-mo)

F,M,T 24 1.0 1.85 – – h No

Alevin, 1-day 

old

F,M,T 24 27 49.90 – – gt No

Smolt (68 g, 

13 mo)

F,M,T 24 2.90 5.36 – – gt No

2.65 g F,M,T 9 0.5 2.06 – – Cusimano and others, 

1986

h, 26 No

Fry, 1.0 g F,M,T 29 2.09 3.30 – – Davies and Brinkman, 

1994b

h Yes

45–50 mm, 

36 g

F,M,T 45 2.64 2.88 – – Davies and others, 1993 h Yes

Juvenile, 

50 mm

F,M,T 50 3.08 3.08 – – h, 18 Yes

Not reported F,M,T 31 1.75 2.61 – – U.S. Environmental 

Protection Agency, 2001, 

citing Davies, 1976

– No

Swimup fry, 

Kootenai 

strain

R,M,D 21 0.40 0.83 – – EVS Environment 

Consultants, 1996; 

Windward, 2002

h, 17 Yes

1.0 g, pH 7.5 F,M,T 28 0.38 0.61 – – Hansen and others, 1999; 

Hansen and others, 2002a

h, 19 Yes

1.1 g, pH 7.5 F,M,T 28 0.47 0.76 – – h, 19 Yes

0.26 g, pH 7.5 F,M,T 29 0.71 1.12 – – h, 19 Yes

0.3 g, pH 6.5 F,M,T 29 1.29 2.03 – – h, 20 Yes

0.66 g, pH 7.5 F,M,T 30 0.51 0.79 – – h, 21 Yes

0.28 g, pH 7.5 F,M,T 83 2.85 1.86 – – h, 19 Yes

Juvenile, ~4.5 g F,M,T 140 22 9.30 – – Hollis and others, 1999 h Yes

Juvenile, ~ 12 g F,M,T 20 0.77 1.66 – – Hollis and others, 2000a h, 22 Yes

Juvenile, ~ 12 g F,M,T 20 0.61 1.31 – – h, 23 Yes

Juvenile, ~ 12 g F,M,T 20 2.07 4.46 – – h, 24 Yes

Juvenile, size 

not reported

F,M,T 25 2.35 4.19 – – Hollis and others, 2000b h, 25 Yes

Juvenile, size 

not reported

F,M,T 47 2.35 2.47 – – h Yes

Juvenile, size 

not reported

F,M,T 77 2.15 1.50 – – h Yes

96 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Oncorhynchus 

mykiss—

Continued

Rainbow 

trout—

Continued

Juvenile, size 

not reported

F,M,T 120 1.15 0.55 – – Hollis and others, 2000b h Yes

Juvenile, 8-12 g F,M,T 120 19 9.13 – – Niyogi and others, 2004 h Yes

8.8 g F,M,T 44 3 3.31 – – Phipps and Holcombe, 

1985

h No

0.5 g S,M,T 44 2.30 2.58 – – Spehar and Carlson, 1984 h No

Juvenile, 18 g F,M,T 50 1.88 1.88 – – Stubblefield, 1990 h Yes

Swimup fry, 

Mt Lassen 

strain, avg. 

wt.(g)= 0.39

R,M,D 7.5 0.48 2.33 – – Windward, 2002 h Yes

Swimup fry, 

Mt Lassen 

strain, avg. 

wt.(g)= 0.39

R,M,D 14 0.97 2.89 – – h Yes

Swimup fry, 

Mt Lassen 

strain

R,M,D 21 0.84 1.74 – – h Yes

Swimup fry, 

Mt Lassen 

strain, avg. 

wt.(g)= 0.39

R,M,D 24 1.30 2.40 – – h Yes

Swimup fry, 

Mt Lassen 

strain, avg. 

wt.(g)= 0.36

R,M,D 29 0.83 1.33 – – h Yes

Swimup fry, 

Mt Lassen 

strain, avg. 

wt.(g)= 0.39

R,M,D 30 0.99 1.51 – – h, 38 Yes

Swimup fry, 

Mt Lassen 

strain, avg. 

wt.(g)= 0.31

R,M,D 32 0.89 1.29 – – h Yes

Swimup fry, 

4–5 weeks 

F,M,T 101 3.80 2.13 – – Besser and others, 2007 h Yes

Swimup fry, 

4–5 weeks

F,M,T 101 5.30 2.97 – – h Yes

Swimup fry, 

4–5 weeks

F,M,T 101 5.40 2.97 2.04 – – Yes

Table 15 97
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Oncorhynchus 

tshawytscha

Chinook salmon Alevin, 0.05 g, 

1-day

F,M,T 24 27 49.90 – – Chapman, 1978b gt No

Swimup 

(0.23 g, 

~3 weeks)

F,M,T 24 1.8 3.33 – – No

Parr (12 g, 

5 months)

F,M,T 24 3.5 6.47 – – 9 No

Smolt (32 g, 

18 months )

F,M,T 24 2.9 5.36 – – gt, 9 No

~ 0.8 g F,M,T 25 1.41 2.52 – – Chapman, 1982 27 No

1.1 g F,M,T 21 1.1 2.27 2.67 2.00 Finlayson and Verrue, 

1982

No

Orconectes 

immunis

Calico crayfish 1.8 g F,M,T 44 10,200 – – – Phipps and Holcombe, 

1985

No

Orconectes 

juvenilis

Crayfish 3rd to 5th 

instars,

~ 0.02 g

R,M,T 44 60 66.8 66.8 – Wigginton, 2005 35 Yes

Orconectes 

placidus

Placid crayfish 3rd to 5th 

instars,

~ 0.02 g

R,M,T 55 37 34.2 34.2 47.8 – Yes

Orconectes virilis Virile crayfish ng F,M,T 26 6,100 10,543 – – U.S. Environmental 

Protection Agency, 2001

– No

Perca flavescens Yellow perch 8–12 g F,M,T 120 8,140 3,912.74 3,912.74 3,913 Niyogi and others, 2004 – Yes

Perlodidae Stonefly Field collected R,M,D 30 5,130 7,865.94 7,866 7,866 EVS Environment 

Consultants, 1996

– Yes

Physa gyrina Physid snail Immature F,M,T 200 410 128.53 128.53 128.5 U.S. Environmental 

Protection Agency, 2001

28 No

Pimephales 

promelas

Fathead 

minnow

4–6 days R,M,T 266 17 4.20 – – Castillo and Longley, 

2001

h Yes

4–6 days R,M,T 278 15.43 3.67 – – h Yes

2–4 days S,M,D 17 4.8 11.84 – – Suedel and others, 1997 h Yes

Fry S,M,T 40 21 25.31 – – U.S. Environmental 

Protection Agency, 2001, 

citing Spehar, 1982

h No

Fry S,M,T 48 11.7 12.11 – – h No

Fry S,M,T 39 19.3 23.76 – – h No

Fry S,M,T 45 42.5 46.42 – – h No

Fry S,M,T 47 54.2 57.08 – – h No

Fry S,M,T 44 29 32.27 – – h No

<24 hr S,M,T 290 65 14.93 – – Schubauer-Berigan and 

others, 1993

h, 8 Yes

0.15 g, 30-days S,M,T 45 13.2 14.42 16.49 16.49 Spehar and Fiandt, 1986 h No

Poecilia reticulata Guppy ng R,M,T 105 3,800 2,042.51 – – U.S. Environmental 

Protection Agency, 2001

– No

ng R,M,T 209 11,100 3,353.90 2,617.33 2,617.33 U.S. Environmental 

Protection Agency, 2001

– No

Potamopyrgus 

antipodarum

New Zealand 

mud snail

Juvenile, 

3–4 mm

R,M,T 148 560 225.9 225.9 225.9 Jensen and Forbes, 2001 – Yes

98 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Procambarus 

clarkii

Red swamp 

crayfish

3rd to 5th 

instars,

~ 0.02 g

R,M,T 42 624 722 722 722 Wigginton, 2005 – Yes

Juvenile S,M,T 30 1,040 1,594.65 – – U.S. Environmental 

Protection Agency, 2001

– No

Prosopium 

williamsoni

Mountain 

whitefish

Field collected, 

209 g

F,M,T 50 8.29 8.29 8.29 8.29 Stubblefield, 1990 gt Yes

Ptychocheilus 

oregonensis

Northern 

pikeminnow

Juvenile F,M,T 25 1,092 1,950.34 1,950.34 1,950.34 U.S. Environmental 

Protection Agency, 2001

– No

Quistadrilus 

multisetosus

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 320 2,092.81 2,092.81 2,092.81 Chapman and others, 

1982

– No

Rhithrogena sp. Mayfly larvae R,M,D 21 50 103.33 103.33 103.33 Windward, 2002 gt Yes

Rhyacodrilus 

montana

Tubificid worm ng S,M,T 5.3 630 4,120.22 4,120.22 4,120.22 Chapman and others, 

1982

– No

Salmo trutta Brown trout Fingerling, 22 g F,M,T 50 2.85 2.85 – – Stubblefield, 1990 h Yes

0.48 g F,M,T 38 2.37 3.01 – – Davies and Brinkman, 

1994c

h Yes

34 days post 

swimup

F,M,T 29 1.23 1.94 – – Brinkman and Hansen, 

2007

h Yes

34 days post 

swimup

F,M,T 68 3.9 3.02 – – h Yes

34 days post 

swimup

F,M,T 151 10.1 4.01 – – h Yes

1-2 g S,M,T 44 1.4 1.57 2.61 2.61 Spehar and Carlson, 1984 h No

Salvelinus 

confluentus

Bull trout 0.07 g, pH 7.5 F,M,T 29 .91 1.43 – – Hansen and others, 1999; 

Hansen and others, 2002a

h, 19 Yes

0.07 g, pH 7.5 F,M,T 83 6.06 3.96 – – h, 19 Yes

0.08 g, pH 6.5 F,M,T 29 2.89 4.54 – – h, 20 Yes

0.2 g, pH 7.5 F,M,T 30 1.0 1.54 – – h, 21 Yes

0.22 g, pH 7.5 F,M,T 28 .90 1.44 – – h, 19 Yes

0.22 g, pH 7.5 F,M,T 28 .99 1.61 2.13 2.13 h, 19 Yes

Simocephalus 

serrulatus

Cladoceran ng S,M,T 11 7 24.85 – – U.S. Environmental 

Protection Agency, 2001

– No

ng S,M,T 44 24.5 27.53 26.15 26.15 Spehar and Carlson, 1984 – No

Spirosperma sp. 

(“ferox”)

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 350 2,943.02 2,943.02 2,943.02 Chapman and others, 

1982

29 No

Spirosperma 

nikolskyi

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 450 2,943.02 2,943.02 2,943.02 Chapman and others, 

1982

– No

Stylodrilus 

heringlianus

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 550 3,597.02 3,597.02 3,597.02 Chapman and others, 

1982

– No

Tubifex tubifex Tubificid worm ng S,M,T 128 3,200 1,457.31 – – U.S. Environmental 

Protection Agency, 2001

h No

ng S,M,T 128 1,700 7,74.20 – – h No

Probably 

adult, field 

collected

R,M,T 5.3 320 2,092.81 1,331.61 1,331.61 Chapman and others, 

1982

h No

Table 15 99
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Species Common 

names

Size or age at 

test initiation Method

Hardness 

(mg/L as 

CaCO3)

LC50 

(μg/L)

Hardness 

adjusted 

LC50 (μg/L)

SMAV 

(μg/L)

GMAV 

(μg/L) Reference Code/ 

Notes

Data 

different 

than EPA 

(2001)?

Utterbackia 

imbecilis

Paper pondshell 

(mussel)

ng S,M,T 90 114.7 70.14 – – U.S. Environmental 

Protection Agency, 2001, 

citing Keller, unpublished

– No

ng S,M,T 90 111.8 68.37 – – – No

Juvenile S,M,T 92 81.9 49.17 – – – No

Juvenile S,M,T 86 93 59.07 – – – No

Juvenile S,M,T 44 9 10.02 – – Keller and Zam, 1991 – Yes

Juvenile S,M,T 90 107 65.43 46.49 46.49 – No

Varichaetadrilus 

pacificus

Tubificid worm Probably 

adult, field 

collected

R,M,T 5.3 380 2,485.21 2,485.21 2,485.21 Chapman and others, 

1982

30 No

Villosa vibex Southern 

rainbow 

(mussel)

Juvenile S,M,T 40 30 36.16 – – U.S. Environmental 

Protection Agency, 2001, 

citing Keller, unpublished

h, 31 No

Juvenile S,M,T 186 125 41.64 38.80 38.80 h, 32 No

Code/Notes: Alpha notes: gt, “greater than” value; h, data used in hardness-toxicity regressions; lt, “less than” value.

Numbered notes: 

1. Listed as “Actinonaia pecorosa” in source.

2. Apparently identical data that were described in more than one publication are assumed to reflect the results of a single test and are only used 

once here.

3. Listed as “Asellus bicrenata” in source. No record found in ITIS (2002).

4. Hardness value obtained by written communication from author.

5. Originally reported as >73 μg/L, but because mortality at 73 μg/L (the highest concentration tested) was 45 percent greater than control 

mortality, this concentration is considered a reasonable estimate of the 50 percent lethality level.

6. Organisms were fed but data considered usable because in comparison to other values for this species, results did not appear to be insensitive 

suggesting feeding did not appear to be an important factor in reducing toxicity. Because aeration was not needed to maintain dissolved 

oxygen levels, organic loading from feeding was not very high (B. Suedell, oral commun.).

7. Total organic carbon concentration in the test water was 7.1 mg/L. Stephan and others (1985) advise against using the results of tests in 

which organic carbon exceeded 5 mg/L in the test water unless a relationship is developed or unless data show that organic carbon does not 

affect toxicity. In comparison to other values for this genus in waters with lower DOC, the responses from this test were about as sensitive. 

Further, in contrast to copper, DOC has been shown to play a minor role in mitigating cadmium toxicity to fish because the binding affinity 

of cadmium to the gill surface is higher than the binding affinity of cadmium to dissolved organic matter (Playle, 2004; Playle and others, 

1993). Thus, it appears that organic carbon in ambient waters at usual concentrations is not an important influence on cadmium toxicity.

8. Only values with natural pH of 8.3 used. Study manipulated pH while holding hardness and alkalinity constant. Only values were used from 

treatments in which pH was not manipulated in a manner conflicting with the assumption that hardness and pH are correlated.

9. Only newly hatched and swimup stage data were used in SMAV calculation because older fish may be more resistant.

10. Value for 72-hour LC50, 72-hour LC50 lower than 48 hours suggesting ACRs based on 48-hour LC50s biased high.

12. Listed as “Glossiponia complanta” in source.

13. Organisms were fed; however, investigations of the influence of feeding on the responses of Hyalella azteca in sediment toxicity testing 

suggest that feeding was unlikely to have biased the results (Ankley and others, 1993).

14. Value estimated from authors’ graph.

15. Sublethal effects, tenacle clubbing and slight contraction of body, noted at lower concentrations than LC50 (LOEC 110 μg/L). Zurich strain, 

male clone. “Hydra vulgaris” commonly referenced in literature, but no record for H. vulgaris found in ITIS (2002).

100 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 15. Acute toxicity of cadmium to freshwater animals.—Continued

Code/Notes: 16. 86 percent dead at 0.5 μg/L (lowest concentration tested), LC50 calculated from original data using U.S. Environmental Protection Agency’s 

(1991) trimmed Spearman-Karber method, adjusted for control mortality, which was 15%. Although ASTM (1997) recommended that an 

acute toxicity test should be considered unacceptable if >10% control mortality occurred, the test was considered acceptable because high 

mortality during the early swim-up fry stage was considered typical for this stock (John Thorpe, IDFG Sandpoint Fish Hatchery Manager, 

Sandpoint, Idaho, oral communication) and the test was the sole data point for an apparently sensitive stock.

17. 72 percent dead at 0.5 μg/L (lowest concentration tested), LC50 calculated from original data using EPA’s trimmed Spearman-Karber 

method.

18. Hardness adjusted from 45 mg/L (ambient) to 50 mg/L with MgSO4. Because adjustment was minimal, the use of Mg-only hardness 

adjustment did not disqualify data.

19. Test conducted at pH 7.5. Value listed was calculated by Probit analysis using responses at 96 hours (Hansen and others, 1999, appendix D).

20. Test conducted at pH 6.5. Value listed was calculated by Probit analysis using responses at 96 hours (Hansen and others, 1999, appendix D).

21. Test conducted at pH 7.5 at 8oC. Value listed was calculated by Probit analysis using responses at 96 hours (Hansen and others, 1999, 

appendix D).

22. Fish pre-exposed for 30 days at 0.07 μg/L, too low to induce Cd acclimation. Fish weights in acute tests were not given directly, estimated 

size of fish used in their acute tests were calculated from specific growth rates obtained during preceding chronic tests.

23. Fish pre-exposed for 30 days at 0.11 μg/L, too low to induce Cd acclimation. Fish weights in acute tests were not given directly, estimated 

size of fish used in their acute tests were calculated from specific growth rates obtained during the preceding chronic tests.

24. Control fish pre-exposed for 30 days at 0.02 μg/L, too low to induce Cd acclimation. Fish weights in acute tests were not given directly, 

estimated size of fish used in their acute tests were calculated from specific growth rates obtained during preceding chronic tests.

25. Data published in paper had a typographic error (reported as 2.53 μg/L on p. 2729). Lydia Hollis, Golder Associates, Calgary, written 

commun., 2006.

26. Author reported value as <0.5 μg/L in table; but here value is estimated as 0.5 μg/L (not less than). Basis is interpretation of authors’ figure, 

as well as descriptions in text and figures of the value being near 0.4 μg/L, the nominal value or test exposure with partial kill. Detection limit 

was 0.5 μg/L.

27. Same cohort as used in chronic test.

28. Adults 3X less sensitive than immature, whereas with Aplexa, adult were more sensitive.

29. Listed as “Spirosperma ferox” in source. No record found in ITIS (2002).

30. “Varichaeta pacifica” in source, changed per ITIS (2002).

31. Listed as “Vilosa vibex” in source.

32. Jackson and others (2000) conducted tests in addition to the two values listed in this table. However, the additional values were not used 

because they were from tests that varied the ionic composition of dilution waters in proportions that seemed unlikely to occur in natural 

waters.

33. Bishop and McIntosh (1981) reported the results obtained with these two acute tests (with identical results) under a heading that seemed to 

indicate they were obtained using fish collected from a cadmium-contaminated lake (their table 2). However, following a careful reading of 

their text, conclude that row heading was missing and in fact the data were obtained from hatchery control fish; thus the data are considered 

valid and usable.

34. Young-of-year (YOY) lake whitefish were less sensitive than yearlings, with 100 percent survival of YOY at 529 μg/L Cd after 96 hours.

35. No species record in ITIS (2002).

36. Tested at 15oC, tests at colder temperature were less sensitive. Additional information on test waters that was not presented in the published 

paper included: alkalinity 250–270 mg/L, hardness, 150–170 mg/L, and pH 8.1–8.4 (Duane Gossiaux, National Oceanic and Atmospheric 

Administration, written commun., 2005).

37. Animals were fed, however LC50s were similar between fed and unfed animals, so data were not rejected from tests with fed animals.

38. Because of non-monotonic concentration responses, the EC50 estimate for this test is more uncertain than others in this test series and was 

estimated using the trimmed Spearman-Karber method.

Table 15 101
Table 16. Chronic toxicity of cadmium to freshwater animals.

[Each row describes one test. Test type: ELS, early life stage test; LC, life cycle test; JGS, juvenile growth and survival test; PLC, partial life cycle test (survival, growth, and reproduction less than one complete life cycle). Method: D, dissolved test concentrations that were measured after being filtered through a 0.45-μm filter; F, flow-through test exposures; M, test concentrations were measured during exposures; R, renewed test exposures; S, static test exposures; T, total metal concentrations determined from unfiltered samples; U, unmeasured test concentrations. NOEC: Highest no observed effect concentrations. LOEC: Lowest observed effect concentration. EC10, EC20: Concentration causing 10 or 20 percent effect in the most sensitive endpoint. Statistic for chronic value: MATC, maximum acceptable threshold concentration. Calculated as the geometric mean of the NOEC and LOEC. SMCV: species mean chronic value. GMCV: genus mean chronic value. Data different than EPA: “No” indicates that the test descriptions and values listed here for a test are identical to those listed in table 2a of U.S. Environmental Protection Agency (2001). “Yes” indicates either that the values listed here either were not included in table 2a of U.S. Environmental Protection Agency (2001), or if they were, some aspect of the test descriptions or summary results in this table differs from those in table 2a of U.S. Environmental Protection Agency (2001). Abbreviations: Cd, cadmium; °C, degrees Celsius; DOC, dissolved organic carbon; EPA, U.S. Environmental Protection Agency; g, gram; ITIS, Integrated Taxonomic Information System; nc, not calculable; μg/L, microgram per liter; μm, micrometer; USGS, U.S. Geological Survey. Symbols: <, less than; >, greater than; –, missing or inapplicable values;~, approximately]

Test 

No. Species Common name Test type Duration (days) Method Lowest endpoint Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

EC10

(μg/L)

EC20

(μg/L)

Statistic 

used for 

chronic 

value 

Chronic 

value 

(CV, μg/L)

Hardness 

adjusted 

CV (μg/L)

SMCV GMCV Code/

notes

Reference

Data 

different 

than EPA 

(2001)?

1 Aelosoma 

headleyi

Polychaete 

worm

LC 14 R,M,T Population 62 17.2 36.9 – – MATC 25.2 22.0 – – – Niederlehner and 

others, 1984

Yes

2 LC 14 R,M,T Population 168 32 50.2 – – MATC 40.1 18.8 – – – Yes3 LC 14 R,M,T Population 189 53.6 92 – – MATC 70.2 30.6 23.3 23.3 Yes4 Ambystoma 

gracile

Northwestern 

salamander

JGS 24 F,M,T Growth 45 12.8 44.6 – – NOEC 23.89 25.5 25.5 25.5 1 Nebeker and others, 

1994, 1995

Yes

5 Aplexa 

hypnorum

Snail LC 26 F,M,T Reproduction 45.3 2.41 4.41 – – MATC 3.26 3.47 – – 2 Holcombe and others, 

1984

Yes

6 LC 26 F,M,T Reproduction 45.3 1.51 2.5 – – MATC 1.94 2.07 2.68 2.68 3 Yes7 Bufo 

americanus

American toad PLC 60 R,M,T Growth 51 <1 5 – – MATC 2.24 2.21 2.21 2.21 4 James and Little, 

2003

Yes

8 Catostomus 

commersoni

White sucker ELS 60 F,M,T Mortality 44 4.2 12 – – MATC 7.10 7.68 7.68 7.68 5 Eaton, 1978 No

9 Ceriodaphnia 

dubia

Water flea LC 7 S,M,D Reproduction 17 1.0 4.0 – – MATC 2.00 3.90 – – 6 Suedel and others, 

1997

Yes

10 LC 7 R,M,T – 20 10 19 – – MATC 13.78 24.3 – – 7 Jop and others, 1995 No11 LC 7 R,M,T – 100 – – – – MATC 2.20 1.43 – – 8 Spehar and Fiandt, 

1986

Yes

12 LC 10 R,M,T Mortality 270 10.7 21.9 nc nc MATC 15.31 5.38 – – 9 Castillo and Longley, 

2001

Yes

13 LC 8 R,M,T Reproduction 261 1.6 3 1.8 2.5 MATC 2.19 .79 – – 10 Yes14 LC 8 R,M,T Reproduction 285 5.7 8.5 5.3 6.4 MATC 6.96 2.37 – – 11 Yes15 LC 8 R,M,T Reproduction 271 9.6 12.2 4.5 6.1 MATC 10.82 3.80 – – 12 Yes16 LC 10 R,M,T Mortality 292 2.8 4 1.7 2.2 MATC 3.35 1.12 – – 13 Yes17 LC 7 R,M,T Reproduction 170 1.1 3.4 – – MATC 1.93 .91 2.04 – – Brooks and others, 

2004

Yes

18 Ceriodaphnia 

reticulata

Water flea LC 9 R,M,T – 44 3.6 7.5 – – MATC 5.20 5.62 5.62 3.39 – Spehar and Carlson, 

1984

Yes

19 Chironomus 

tentans

Midge LC 60 F,M,T Hatching 280 5.8 16.4 – 4.0 MATC 9.75 3.35 3.35 – – Ingersoll and 

Kemble, 2001

No

102 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 16. Chronic toxicity of cadmium to freshwater animals.—Continued

Test 

No. Species Common name Test type Duration (days) Method Lowest endpoint Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) EC10 (μg/L) EC20 (μg/L) Statistic used for chronic value Chronic value (CV, μg/L) Hardness adjusted CV (μg/L) SMCV GMCV Code/ notes Reference Data different than EPA (2001)?

20 Cottus bairdi Mottled sculpin JGS 28 F,M,D Mortality 100 1.4 2.6 – – MATC 1.91 1.24 – – 14 Besser and others, 

2007

Yes

21 JGS 28 F,M,D Mortality 100 .59 1.25 – – MATC .88 .57 0.84 0.84 15 Yes22 Daphnia 

magna

Water flea LC 21 R,M,T Growth 170 .6 2.0 – – MATC 1.1 .51 – – – Baird and others, 

1990

Yes

23 LC 21 R,M,T – 53 0.08 0.29 – 0.07 MATC .15 0.15 – – 16,h Chapman and 

others, 1980; U.S. Environmental Protection Agency, 2001

No

24 LC 21 R,M,T – 103 .16 .28 – .23 MATC .21 .14 – – 16,h Chapman and 

others, 1980; U.S. Environmental Protection Agency,2001

No

25 LC 21 R,M,T – 209 .21 .91 – .33 MATC .44 .18 – – 16,h No26 LC 20 R,M,T Mortality 200 .37 .48 – .37 EC20 .37 .16 – – 17 Canton and Sloof, 

1982

Yes

27 LC 21 R,M,T Reproduction 52 1.97 3.43 1.56 2.14 MATC 2.60 2.54 – – – Chadwick Ecological 

Consultants, 2004a

Yes

28 LC 21 R,M,T Reproduction 99 1.67 3.43 1.89 2.21 MATC 2.39 1.57 – – – Yes29 LC 21 R,M,T Reproduction 150 5 10 – MATC 7.07 3.58 – – – Bodar and others, 

1988

No

30 LC 21 R,M,T Reproduction 130 .22 1.86 1.13 1.65 LOEC 1.65 .91 .54 – 18, 19 Borgmann and others, 

1989a

No

31 LC 21 R,M,T Reproduction 130 .22 1.86 1.65 1.69 LOEC 1.69 .93 – – 19 Borgmann and others, 

1989b

Yes

32 Daphnia pulex Water flea LC 18 R,M,T Reproduction 52 14.6 >14.6 1.5 2.17 EC20 2.17 2.12 – – 20 Chadwick Ecological 

Consultants, 2004a

Yes

33 LC – R,M,T Reproduction 65 – – – – MATC 7.49 6.37 – – – U.S. Environmental 

Protection Agency, 2001

No

34 LC 58 R,M,T Reproduction 106 5 10 – – MATC 7.07 4.44 – – h Ingersoll and Winner, 

1982

Yes

35 LC 42 R,M,T Reproduction 58 2.7 4.8 – – MATC 3.60 3.28 – – h Winner, 1986 Yes36 LC 42 R,M,T Reproduction 116 5.6 10.8 – – MATC 7.78 4.62 – – h Yes37 LC 42 R,M,T Reproduction 230 5.2 10.4 – – MATC 7.35 2.86 3.71 1.44 h Yes38 Esox lucius Northern pike ELS 60 F,M,T Mortality 44 4.2 12.9 – – MATC 7.36 7.97 7.97 7.97 5 Eaton, 1978 No

Table 16 103
Test 

No. Species Common name Test type Duration (days) Method Lowest endpoint Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

EC10

(μg/L)

EC20

(μg/L)

Statistic 

used for 

chronic 

value 

Chronic 

value 

(CV, μg/L)

Hardness 

adjusted 

CV (μg/L)

SMCV GMCV Code/

notes

Reference

Data 

different 

than EPA 

(2001)?

39 Etheostoma 

fonticola

Fountain darter JGS 7 R,M,T Mortality 270 1.4 2.8 3.1 4.2 MATC 1.98 0.70 – – 9 Castillo and Longley, 

2001

Yes

40 JGS 7 R,M,T Mortality 261 5.5 11.5 nc nc MATC 7.95 2.86 – – 10 Castillo and Longley, 

2001

Yes

41 JGS 7 R,M,T Mortality 285 5.7 8.5 3.7 4.6 MATC 6.96 2.37 – – 11 Yes42 JGS 7 R,M,T Mortality 270 6.6 9.6 7.1 8.5 MATC 7.96 2.8 – – 12 Yes43 JGS 7 R,M,T Mortality 292 4 5.3 4.1 4.5 MATC 4.60 1.54 1.83 1.83 13 Yes44 Gammarus 

fasciatus

Amphipod JGS 42 R,M,T Mortality 130 1.49 2.23 1.2 1.5 MATC 1.82 1.01 1.01 1.01 21 Borgmann and others, 

1989b

Yes

45 Hyalella azteca Amphipod JGS 14 S,M,D Mortality 17 0.25 – – MATC .16 .31 – – 6,h Suedel and others, 

1997

Yes

46 JGS 28 F,M,T Mortality 280 .51 1.9 1.1 1.40 MATC .98 .34 – – h Ingersoll and 

Kemble, 2001

No

47 JGS 28 R,M,T Mortality 126 .5 1.1 .30 .50 MATC .74 .42 – – h Chadwick Ecological 

Consultants, 2004a

Yes

48 JGS 28 R,M,T Mortality 153 .8 1.3 .40 .70 MATC 1.02 .51 – – h Yes49 JGS 42 R,M,T Mortality 130 .18 .34 MATC .25 .14 – – – Borgmann and others, 

1991

Yes

50 JGS 42 R,M,T Mortality 130 .57 .92 .64 .81 MATC .72 .40 – – h Borgmann and others, 

1989b

Yes

51 JGS 42 R,M,T Mortality 163 .48 .94 – – MATC .67 .32 .33 .33 22,h Stanley and others, 2005 Yes52 Jordanella 

floridae

American 

flagfish

LC ~100 F,M,T Reproduction 44 4.1 8.1 4.5 5.3 MATC 5.76 6.24 – – – Spehar, 1976 Yes

53 LC – – – 47.5 3 6.5 – – MATC 4.42 4.56 – – – U.S. Environmental 

Protection Agency, 2001

No

54 LC – – – 47.5 3.4 7.3 – – MATC 4.98 5.14 5.27 5.27 – No55 Lepomis 

macrochirus

Bluegill LC – – – 207 31 80 – – MATC 49.80 20.6 20.6 – – U.S. Environmental 

Protection Agency, 2001

No

56 Micropterus 

dolomieui

Smallmouth bass ELS 60 F,M,T Mortality 44 4.3 12.7 – – MATC 7.39 8.00 8.00 – 5 Eaton, 1978 No

57 Oncorhynchus 

kisutch

Coho salmon ELS 27 F,M,T Mortality 44 1.3 3.4 – – MATC 2.10 2.28 2.28 – 23 Eaton, 1978 No

Table 16. Chronic toxicity of cadmium to freshwater animals.—Continued

104 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Test 

No. Species Common name Test type Duration (days) Method Lowest endpoint Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) EC10 (μg/L) EC20 (μg/L) Statistic used for chronic value Chronic value (CV, μg/L) Hardness adjusted CV (μg/L) SMCV GMCV Code/ notes Reference Data different than EPA (2001)?

58 Oncorhynchus 

mykiss

Rainbow trout ELS 53 F,M,D Mortality 20 0.6 1.3 0.89 1.21 MATC 0.88 1.56 – – h Mebane and others, 

2008

Yes

59 ELS 62 F,M,D Mortality 29 1.0 2.5 1.61 2.21 MATC 1.58 2.22 – – h Yes60 ELS 665 F,M,T Reproduction 250 3.39 5.48 – – MATC 4.31 1.59 – – 24,h Brown and others, 

1994

No

61 ELS 100 F,M,T Mortality 46 1.25 1.74 2 – MATC 1.47 1.55 – – 25,h Davies and others, 

1993

Yes

62 JGS 100 F,M,T Mortality 29 1.02 1.89 – – MATC 1.39 1.95 – – h Davies and 

Brinkman, 1994b

Yes

63 JGS 100 F,M,T Mortality 282 2.24 7.02 – – MATC 3.97 1.36 1.68 – 26,h Yes64 Oncorhynchus 

tshawytscha

Chinook salmon ELS 120 F,M,T Mortality 25 .96 1.3 1.33 1.83 MATC 1.12 1.72 1.72 1.87 27 Chapman, 1975, 1982 Yes

65 Oreochromis 

aurea

Blue tilapia LC – – – 145 52 – – – NOEC 52.00 26.9 26.9 26.9 – U.S. Environmental 

Protection Agency, 2001

No

66 Pimephales 

promelas

Fathead minnow JGS 10 S,M,D Mortality 17 1 2 – – MATC 1.41 2.76 – – 6 Suedel and others, 

1997

Yes

67 JGS 14 S,M,D Mortality 17 2 3 – – MATC 2.45 4.78 – – – Yes68 ELS 32 F,M,T Mortality 44 – – – – MATC 10.00 10.8 – – – Spehar and Fiandt, 

1986

No

69 ELS 32 F,M,T Mortality 44 9 18 – – MATC 12.73 13.8 – – – Spehar and Carlson, 

1984

Yes

70 LC ~300 F,M,T Mortality 201 37 110 – – MATC 63.80 26.8 – – – Pickering and Gast, 

1972

No

71 LC ~250 F,M,T Mortality 201 27 57 – – MATC 39.23 16.5 – – – No72 JGS 7 R,M,T Mortality 270 10.7 21.9 7.0 12.8 MATC 15.31 5.38 – – 9 Castillo and Longley, 

2001

Yes

73 JGS 7 R,M,T Mortality 261 11.5 21.3 nc – MATC 15.65 5.62 – – 10 Yes74 JGS 7 R,M,T Mortality 285 8.5 11.3 11.3 12.6 MATC 9.80 3.33 – – 11 Yes75 JGS 7 R,M,T Mortality 272 9.6 12.2 8.7 10.8 MATC 10.82 3.79 – – 12 Yes76 JGS 7 R,M,T Mortality 292 5.3 6.9 5.9 7.3 MATC 6.05 2.03 6.35 5.9 13 Yes77 Salmo salar Atlantic salmon ELS 92 F,M,T Alevin 

biomass

28 .47 .78 – – MATC .61 .87 .87 .87 28 Rombough and 

Garside, 1982

Yes

Table 16. Chronic toxicity of cadmium to freshwater animals.—Continued

Table 16 105
Table 16. Chronic toxicity of cadmium to freshwater animals.—Continued

Test 

No. Species Common name Test type Duration (days) Method Lowest endpoint Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

EC10

(μg/L)

EC20

(μg/L)

Statistic 

used for 

chronic 

value 

Chronic 

value 

(CV, μg/L)

Hardness 

adjusted 

CV (μg/L)

SMCV GMCV Code/

notes

Reference

Data 

different 

than EPA 

(2001)?

78 Salmo trutta Brown trout JGS 84 F,M,T Mortality 37.6 – 0.7 – – LOEC 0.70 0.84 – – 29 Davies and 

Brinkman, 1994c

Yes

79 JGS 136 F,M,T Mortality 39.6 1 1.78 – – MATC 1.33 1.54 – – – Davies and 

Brinkman, 1994a

Yes

80 ELS 55 F,M,T Mortality 31 2.54 4.87 – – MATC 3.52 4.73 – – h Brinkman and 

Hansen, 2007

Yes

81 JGS 30 F,M,T Mortality 29 0.74 1.4 – – MATC 1.02 1.43 – – – Yes82 ELS 55 F,M,T Mortality 71 4.68 8.64 – – MATC 6.36 5.12 – – h Yes83 JGS 30 F,M,T Mortality 68 1.3 2.58 – – MATC 1.83 1.51 – – – Yes84 ELS 55 F,M,T Mortality 149 9.62 19.1 – – MATC 13.56 6.89 – – h Yes85 JGS 30 F,M,T Mortality 151 4.81 8.88 – – MATC 6.54 3.29 – – – Yes86 LC 665 F,M,T Mortality 250 9.3 29.1 – – MATC 16.45 6.07 – – h Brown and others, 

1994

No

87 ELS 60 F,M,T Mortality 44 3.8 11.7 MATC 6.67 7.22 3.04 1.62 5,h Eaton, 1978 No88 Salvelinus 

confluentus

Bull trout JGS 55 F,M,T Mortality 29 .38 .78 0.49 0.64 MATC .55 .77 .77 – – Hansen and others, 

2002b

Yes

89 Salvelinus 

fontinalis

Brook trout LC 1,100 F,M,T Mortality 44 1.7 3.4 – – NOEC 1.70 1.84 – – 30 Benoit and others, 

1976

Yes

90 ELS 60 F,M,T Mortality 37 1 3 1.7 2.8 MATC 1.70 2.05 – – – Sauter and others, 

1976

No

91 ELS 60 F,M,T Mortality 188 3 7 2.9 7.9 MATC 4.58 2.02 – – – Yes92 ELS 60 F,M,T Mortality 44 1.1 3.8 – – MATC 2.00 2.16 2.12 1.28 5 Eaton, 1978 No93 Salvelinus 

namaycush

Lake trout ELS 60 F,M,T Mortality 44 4.4 12.3 – – MATC 7.36 7.96 7.96 – 5, 31 Eaton, 1978 No

Code/Notes: Alpha notes: h, data used in hardness-toxicity regressions.

Numbered notes: 

1 Test considered acceptable for the chronic dataset because exposure durations and measured endpoints were similar to those used for chronic testing with fish (fish and salamanders have 

somewhat similar life spans), and because the EPA authors stated that the impetus for the testing the effects of long-term Cd exposures to salamanders was to address a data gap in the aquatic life criteria dataset. Unaware of any guidance on test durations/endpoints sufficient to elicit sensitive, chronic responses in salamanders. Leg regeneration was the most sensitive endpoint reported, with significantly slower regrowth at LOEC than in controls.

2 4.4 μg/L treatment considered LOEC: Eggs surviving to test termination was 57 percent of control. Statistical power very low because only 2 replicates (Holcombe and others, 1984, table 1). 3 2.5 μg/L treatment considered LOEC: <50 percent of control hatchability, <50 percent embryo survival, and 30 percent of control egg production (Holcombe and others, 1984, table 1). 4 NOEC was control, <1.0 μg/L. Cd caused 5-6 percent accelerated growth and metamorphosis at both 5 and 54 μg/L treatments, considered an adverse stress response by the authors. 

Decreased survival only observed at highest concentration tested (540 μg/L). Chronic value is uncertain because only three treatments were used with 10 times difference in concentrations (5, 54, and 540 μg/L) and because effects were observed at the lowest concentration tested. MATC value using less than detection control concentrations is greatly dependent on value assumed, assuming lower values such as 0.05 or 0.5 (i.e., half the detection value from other studies at same lab which used more sensitive methods, or half the detection limit from this study) would result in MATC much lower than lowest observed effect concentration. Because magnitude of observed effects seemed fairly small, such extrapolation does not seem justified by the information presented. 

5 No information on magnitude of response given.

106 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 16. Chronic toxicity of cadmium to freshwater animals.—Continued

Code/Notes: 6 Tests used static technique, however data considered usable because results appear reasonable in comparison to other tests with this species. Decision to use static testing in this test 

series was influenced by analyses showing that exposure concentrations, pH, and dissolved oxygen concentrations were acceptably stable. Organic loading from feeding was minimal, as evidenced by the lack of need for aeration (B. Suedell, Entrix, Inc., oral commun., 2005).

7 This chronic value was 31 times higher than the lowest hardness-adjusted chronic value for this species. Value was considered anomalous and was not used in GMCV calculations, even 

though the test appeared otherwise acceptable. 

8 St. Louis river water with DOC >5 mg/L, but low values and a ratio of 1.0 for chronic C. reticulata value with reconstituted lab water and St. Louis water value suggest minimal DOC 

influence. See also table 16, note 7.

9 Test initiated 11/12/1999. Hardness values were not included in the source for the C. dubia test series, but were the same as hardness values from concurrent fountain darter and fathead 

tests that all used a common batch of dilution water (V. Castillo, Southwest Texas State University, written commun., 2005).

10 Test initiated 12/3/1999.11 Test initiated 1/28/2000.12 Test initiated 2/4/2000.13 Test initiated 2/27/2000.14 2003 test using Minnesota strain.15 2004 test using Missouri strain.16 EC20 values from U.S. Environmental Protection Agency, 2001. Source of U.S. Environmental Protection Agency’s (2001) values is assumed to have been an unattributed author 

communication, because neither the EC20 values nor sufficient information to calculate them, were presented in the attributed source.

17 Effects were observed at lowest test concentration but magnitude of effect not given. The EC20 was considered the best NOEC estimate.18 LOEC=20 percent reduction in brood size. Considered better estimate of chronic value than listing as “less than” or as the geometric mean of the control (0.22 μg/L NOEC) and LOEC. 

LOEC was lowest treatment.

19 Similar data that were described in more than one publication are assumed to reflect the results of a single test and are only one value is used.20 MATC >14.6 μg/L, IC20 used instead as chronic value, assuming a 20 percent reduction in survival is at least biologically significant if not statistically significant.21 NOEC resulted in 30 percent reduction in growth from controls even though not statistically different from control. Authors considered 55 percent mortality to be the LOEC and 18 percent 

mortality and 30 percent growth reduction to be the NOEC. Were the 30 percent growth reduction to have been considered the LOEC the MATC would have been 1.17 μg/L (0.92 and 1.49 μg/L limits). Control survival was only 45 percent. Because a concentration response was observed with cadmium, and controls and low treatment concentrations were statistically distinguished from higher treatments by analysis of variance (ANOVA), the responses were considered acceptably reliable and the data usable.

22 Tested in reconstituted hard water. In matched tests conducted in effluent dominated stream mesocosms or laboratory tests using effluent dominated dilution water, cadmium was 

considerably less toxic.

23 Test initiated with sac fry about 1 week after hatching.24 Lowest effect was with second generation fry development, which was significantly affected when the parents were exposed to 1.77 μg/L cadmium, but not when exposed to 0.47 μg/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. The ability of the first generation adults to reach sexual maturity, with NOEC and LOEC values of 3.39 and 5.48 μg/L cadmium, respectively was considered a more reliable endpoint. See also U.S. Environmental Protection Agency (2001, p. 15)

25 Long term growth and survival test, initiated with 36 g juveniles.26 Authors considered 2.5 percent mortality LOEC with no statistical comparisons. 2.5 percent mortality is considered here to be a biologically minor effect, and is considered the NOEC 

concentration. 7.3 μg/L total Cd considered LOEC with 37.5 percent mortality. Test conducted in “aged” hard water (allowed to equilibrate). 

27 190 percent increase in mortality over controls (18 vs... 9 percent) with at 1.3 μg/L treatment, which is considered the LOEC.

Table 16 107
Table 16. Chronic toxicity of cadmium to freshwater animals.—Continued

Code/Notes: 28 Statistical LOEC of 0.47 μg/L resulted in reduced alevin weight and growth, lowest concentration tested, although authors suggest threshold would be 0.13 μg/L (p. 2012) - “We observed 

significant reduction in the growth of salmon alevins in 0.47 μg/L and examination of a plot of final alevin wet weight versus ambient cadmium concentration indicated a response threshold near the control concentration of 0.13 μg/L. This suggests that levels currently assumed to be safe may have some deleterious effect on populations, although probably of such a small magnitude as to be of little consequence.” Based on this statement, conclude statistical LOEC is more like a biological NOEC. Next higher concentration also had reduced survival and statistically significant reduction in biomass (their table 6). To summarize effects for this synthesis, treatment 1 (0.47 μg/L) is considered the NOEC with biomass 99 percent of control, treatment 2 (0.78 μg/L) is considered LOEC with biomass 72 percent of control. Average zinc in control water was 95 μg/L during 1977 test and 80 μg/L during the 1978 test. Because response pattern tracked cadmium concentrations with adequate control survival, and responses were relative to controls, background zinc probably did not compromise tests. One each tests were conducted in the winters of 1977 and 1978. Results from 1977 were much more resistant than results from 1978, only the 1978 results were used for this report. Hardnesses were 19 mg/L (Ca:Mg ratio 2:1) in winter 1977 and 28 mg/L (Ca:Mg ratio 7:1) in winter 1978. The more sensitive results were obtained in 1978 with average temperature of 10°C and pH 7.3 compared with average temperature of 5°C and pH 6.5 in 1977.

29 25 percent mortality at lowest concentration tested30 NOEC judged to be best estimate of chronic value. MATC considered neither acceptable nor a threshold because the LOEC resulted in complete mortality to spawning males.31 This chronic value was >10 times higher than the lowest hardness-adjusted chronic value for this species. Value was considered anomalous and was not used in GMCV calculations, even 

though the test was presumably otherwise acceptable. 

108 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Table 17. Other data on effects of cadmium to freshwater organisms.

[Each row describes one test. Underlined values are lower than their corresponding acute or chronic criterion values. Size or age at test initiation: CSL, chronic sublethal; ELS, early life stage; PLC, partial life cycle (survival, growth, and reproduction, less than one complete life cycle). Method: D, dissolved test concentrations measured after being filtered through a 0.45-μm filter; F, flow-through test exposures; M, test concentrations measured during exposures; R, renewed test exposures; S, static test exposures; T, total metal concentrations determined from unfiltered samples; U, unmeasured test concentrations; ?, unknown. NOEC: Highest no observed effect concentrations. LOEC: Lowest observed effect concentration. Summary statistic: MATC, maximum acceptable threshold concentration. Calculated as the geometric mean of the NOEC and LOEC. LC50, concentration killing 50 percent of test organisms; ILL, incipient lethal level. Notes: ALC, aquatic live criteria; Cd, cadmium; Cu, copper; Hg, mercury; MgSO4, magnesium sulfate. Abbreviations: °C, degrees Celsius; DOC, dissolved organic carbon; EPA, U.S. Environmental Protection Agency; g, gram; ITIS, Integrated Taxonomic Information System; g, gram; GMCV, genus mean chronic value; nc, not calculable; μg/L, microgram per liter; μm, micrometer; mg/kg, milligram per kilogram; mm, millimeter; Pb, lead; SMAV, species mean acute value; SMCV, species mean chronic value; USGS, U.S. Geological Survey; YOY, young-of-year fish; <, less than; >, greater than; ~, approximately; –, missing or inapplicable values]

Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Ambystoma 

gracile

Northwestern 

salamander

3-mo old 

larvae

24 F,M,T Growth 45 48.9 193 MATC 97.15 0.38 Overall growth and survival were 

less sensitive endpoints than limb regeneration.

Nebeker and others, 

1994, 1995

Brachionus 

calyciflorus

Rotifer Cysts 2 S,U Reproduction 90 18 25 MATC 21.2 14.7 Species has a cyclically 

parthenogenetic life cycle; Sexual reproduction (mictric phase) more affected by contaminants than asexual (amictic phase).

Snell and Carmona, 

1995

Brachycentrus 

sp.

Caddisfly 5-8 mm 28 F,M,T Behavior 45 27.5 85.5 MATC 48.5 .36 Behavior of caddisflies at 85 and 

238 μg/L was abnormal. The affected larvae showed a vigorous curling motion and were usually free swimming. Essentially all control larvae were observed in larval cases and responded slowly to probing.

Spehar and others, 

1978a

Bufo 

americanus

American 

toad

Tadpoles 103 M,S,T Mortality and 

growth

60 – .5 LOEC .45 .43 Slightly reduced survival (80 vs.. 

88 percent for controls), delayed development of tadpoles, and a non-significant trend of older and smaller metamorphs exposed to Cd exposed to Cd through periphyton (about 10 mg/kg dry weight and water (5 to <0.1 μg/L). Adverse effects were noted in the lowest cadmium exposure which was dosed with a single slug resulting in an initial nominal aqueous concentration of 5 μg/L. Aqueous concentrations dropped to <0.1 μg/L during the course of the test. Tissue-residue LOEC was 16.7 mg/kg dry weight (whole body).

James and others, 

2005

Table 17 109
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Ceriodaphnia 

dubia

Water flea Neonate 7 R,U Reproduction 90 0.5 1.0 MATC 0.71 0.55 Reproductive endpoints more 

sensitive than survival, and were as low as those generated from longer duration tests. No analytical confirmation of exposure concentrations.

Winner, 1988

Chironomus 

riparius

Midge 10-12 mm 

(late instar)

4 F,M,T Mortality 152 – – LC50 300,000 1.84 Late instar insects may be much less 

sensitive than early instars (1 and 2).

U.S. 

Environmental Protection Agency, 2001

Coregonus 

clupeaformis

Lake whitefish 10 month 

old

.05 F,M,T Behavior 81 <.2 .2 LOEC .2 .51 Detected and responded to area of 

trough with Cd; response could be either preference or avoidance.

McNicol and 

Scherer, 1991

Coregonus 

clupeaformis

Lake whitefish 10 month 

old

.1 F,M,T Behavior 81 1 5 MATC 2.2 .51 Avoidance under uniform lighting 

conditions.

McNicol and 

others, 1999

Coregonus 

clupeaformis

Lake whitefish 10 month 

old

.75 F,M,T Behavior 81 50 125 MATC 79 .51 Cd began to force fish to leave shaded 

area of the tank, which they were strongly attracted to.

McNicol and 

others, 1999

Coregonus 

clupeaformis

Lake whitefish 10 month 

old

14 F,M,T Mortality 81 – – LC50 361 1.1 Young-of-year fish (YOY) less 

sensitive than yearlings.

McNicol, 1997; 

McNicol and others, 1999

Crangonyx 

pseudogracilis

Amphipod Adult, 4 

mm

4 R,U Mortality 50 – – LC50 1,700 .38 Test value indicates that the order 

Amphipoda is not uniformly sensitive to Cd. Test not used in acute dataset because of no analytical confirmation of exposure concentrations.

Martin and 

Holdich, 1986

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

110 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Daphnia galeata 

mendota

Water flea Mixed 

population

154 R,M,T Population 136 1 2 MATC 1.4 0.71 Reduction in carrying capacity. Due 

to compensatory increases, population attributes were less sensitive than most sensitive individual attributes. Study was located too late to include in chronic criterion dataset, although it would have been otherwise usable. Water hardness not reported, hardness values for the same water source ranged from 135-137 mg/L in 2003-05, city of Chicago tapwater (filtered Lake Michigan water), www.cityofchicago.org/water

Marshall, 1978

Daphnia magna Water flea £24hrs old 4 S,M,T Mortality 17 – – LC50 12.7 .31 Data point not used because most other 

data for Daphnia magna were from 48-hour exposures with no feeding. Feeding may reduce cadmium toxicity in 48-hour exposures by about 2 times (Lewis and Weber 1985). However, Suedel and others (1997) data shows that Daphnia were about 2 times more sensitive in 96-hour exposures with feeding than in 48-hour exposures without feeding.

Suedel and 

others, 1997

Daphnia magna Water flea £24hrs old 2 S,M,T Mortality 17 – – LC50 26.4 .31 – Suedel and 

others, 1997

Daphnia magna Water flea Neonate 7 R,U Growth 90 1 2 MATC 1.4 .55 Concluded that a 7-day test was 

insufficiently sensitive for D. magna reproduction endpoints, and recommended C. dubia as a reasonably sensitive shortterm chronic test. No analytical confirmation of exposure concentrations.

Winner, 1988

Daphnia magna Water flea 8-days old 8 R,M,T Reproduction 170 .2 .4 MATC .28 .81 Two instars were exposed starting at 

8-days. 

Barata and 

Baird, 2000

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 111
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Ephemerella sp. Mayfly 5-8 mm 28 F,M,T Mortality 45 – 3 MATC 1.50 0.36 LC50 ~3 μg/L, adjusted for 31 

percent control mortality, higher treatments resulted in 84-94 percent mortality although not consistent with concentration. Almost all control and treatment mortality occurred after day 21. Treatment spacing and lack of concentration-response precludes calculating statistics, LC50/2 is MATC best guess.

Spehar and others, 

1978a

Gila elegans Bonytail chub Larva 4 S,U Mortality 199 – – LC50 148 2.30 No analytical confirmation of 

exposure concentrations.

Buhl, 1997

Gila elegans Bonytail chub Juvenile 4 S,U Mortality 199 – – LC50 168 2.30 No analytical confirmation of 

exposure concentrations.

Buhl, 1997

Hyalella azteca Amphipod Adults and 

larger juveniles

4 S,M,T Mortality 34 – – LC50 8 .54 This acute test included adult 

organisms and also was the highest acute value for the species (~10 times greater than lowest value). While no agesensitivity relationship is known of for this species, it still seemed possible that the most sensitive life stage was not tested and thus the value was excluded from the SMAV.

Nebeker and others, 

1986b

Hyalella azteca Amphipod 7-14 days old 10 F,M,T Mortality 45 – – LC50 2.8 .68 Hardness assumed from other tests 

at EPA’s Duluth lab using Lake Superior water.

Phipps and others, 

1995

Hyalella azteca Amphipod Juvenile 10 F,M,T Mortality 45 – – LC50 2.8 .68 Unclear if the reported results 

reflected two tests with identical results or the same test described twice. 

Ankley and others, 

1991

Hyalella azteca Amphipod 1-11 days 7 S,M,T Mortality 18 – – LC50 .15 .20 Cd was the most toxic of the 63 

metals tested with Hyalella at low hardness. Seven-day test is still an acute test, but was a better predictor of chronic toxicity than were 48- or 96-hour tests. Here the 7-day LC50s are compared to the chronic criterion values.

Borgmann and 

others, 2005

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

112 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Hyalella azteca Amphipod 1-11 days 7 S,M,T Mortality 124 – – LC50 1.6 0.67 Cd was the most toxic of the 63 

metals tested with Hyalella at moderate hardness. Sevenday test is still an acute test, but was a better predictor of chronic toxicity than were 48- or 96-hour tests. Here the 7-day LC50s are compared to the chronic criterion values.

Borgmann and others, 

2005

Hyalella azteca Amphipod 2-3 week old 

juveniles

3 R,M,T Mortality 8 LC50 1.9 .17 Non-standard duration. Gust, 2006

Hydropsyche 

betteni

Caddisfly 5-8 mm 28 F,M,T Behavior 45 27.5 85.5 MATC 48.49 .36 Behavior of caddisflies at 85 

and 238 μg/L was abnormal. The affected larvae showed a vigorous curling motion and were usually free swimming. Essentially all control larvae were observed in larval cases and responded slowly to probing

Spehar and others, 

1978a

Lumbriculus 

variegatus

Oligochaete Mixed age 10 F,M,T Mortality 45 – – LC50 158 .68 10- and 4-day LC50s were similar. Ankley and others, 

1991

Lymnaea 

palustris

Pulmonate 

pond snail

PLC ~100 R,U Hatching 

(embryogenesis inhibition)

250 <40 40 LOEC 40 1.03 Complete inhibition of hatching. 

Cd concentrations unmeasured, however, in other studies in their lab they always had nominal Cd within 30 percent of measured in 1-week static exposures. (M. Coeurdassier, University of Franche-Comte, Besançon, France, written commun., 2005)

Coeurdassier and 

others, 2003

Lymnaea 

stagnali

Pulmonate 

pond snail

4-week old 

juvenile

4 S,M,T Mortality 250 – – LC50 742 2.77 Juveniles apparently more 

sensitive than older animals.

Coeurdassier and 

others, 2004

Lymnaea 

stagnali

Pulmonate 

pond snail

9-week old 

juvenile

4 S,M,T Mortality 250 – – LC50 1,515 2.77 Juveniles apparently more 

sensitive than older animals.

Coeurdassier and 

others, 2004

Lymnaea 

stagnali

Pulmonate 

pond snail

20-week old 

adult

4 S,M,T Mortality 250 – – LC50 1,585 2.77 Juveniles apparently more 

sensitive than older animals.

Coeurdassier and 

others, 2004

Lymnaea 

stagnali

Pulmonate 

pond snail

PLC 49 S,U Hatching 250 <25 25 – 25 1.03 No analytical confirmation of 

exposure concentrations.

Gomot, 1998

Moina 

macrocopa

Planktonic 

cladoceran

24 hrs old 4 S,U Mortality 396 LC50 680 4.03 No analytical confirmation of 

exposure concentrations.

García and others, 

2004

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 113
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Moina 

macrocopa

Planktonic 

cladoceran

24 hrs old 15 R,U Population 

growth rate

396 5 10 MATC 7.07 1.36 No analytical confirmation of 

exposure concentrations.

García and others, 2004

Moina 

macrocopa

Planktonic 

cladoceran

20 R,U Reduced 

survival

82 .2 0.4 MATC .28 .52 No analytical confirmation of 

exposure concentrations.

Hatakeyama and 

Yasuno, 1981

Moina 

triserialis

Planktonic 

cladoceran

24 hrs old 4 S,U Mortality 396 – – LC50 420 4.03 No analytical confirmation of 

exposure concentrations.

García and others, 2004

Moina 

triserialis

Planktonic 

cladoceran

24 hrs old 31 R,U Population 

growth rate

396 11 21 MATC 15.2 1.36 No analytical confirmation of 

exposure concentrations.

García and others, 2004

Morone 

saxatilis

Striped bass 63-days old 4 S,U Mortality 40 – – LC50 4 .62 No analytical confirmation of 

exposure concentrations.

Palawski and others, 

1985

Morone 

saxatilis

Striped bass 63-days old 4 S,U Mortality 285 – – LC50 10 3.08 No analytical confirmation of 

exposure concentrations.

Palawski and others, 

1985

Oncorhynchus 

kisutch

Coho salmon Embryo 

(ELS)

82 F,M,T Mortality 45.3 4.1 12.5 MATC 7.16 .36 Exposed as eyed eggs 20 days 

before hatching which may have induced acclimation. Fish were 4 times more resistant than companion test with coho salmon which introduced fish 1-week after hatching. The two tests used fish from different stocks, so whether the differences are due to acclimation or stock differences is unknown. Since the differences could have been from acclimation influences, the more resistant results were not used in criteria calculations.

Eaton, 1978

Oncorhynchus 

kisutch

Coho salmon Adult 4 F,M,T Mortality 22 – – LC50 17.5 .38 Only the most sensitive life stages 

(swim-up fry and parr) used directly to estimate acute values. More resistant younger and older life stages not used in acute criteria calculations.

Chapman, 1975

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

114 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Oncorhynchus 

kisutch

Coho salmon 2-weeks old 4 R,M,T Mortality 22 – – LC50 1,000 0.38 LC50 “greater than” value. Cu, Pb, 

and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-”clean” methods and unlikely to contribute to test mortalities. Control mortalities not reported, assumed to be zero.

Chapman, 1975

Oncorhynchus 

kisutch

Coho salmon 4-weeks old 4 R,M,T Mortality 22 – – LC50 3.66 .38 LC50 calculated by trimmed 

Spearman-Karber method (“TSK,” U.S. Environmental Protection Agency, 1992) using nominal concentrations (exposures were reported to have been measured but data were not presented). Control mortalities not reported, assumed to be zero. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-”clean” methods. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-”clean” methods, unlikely to contribute to test mortalities.

Chapman, 1975

Oncorhynchus 

kisutch

Coho salmon 5-weeks old 4 R,M,T Mortality 22 – – LC50 2.76 .38 LC50 calculated using Probit 

regression (U.S. Environmental Protection Agency, 1992). Exposures reported to be measured but data not presented. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-”clean” methods. Control mortalities not reported, assumed to be zero.

Chapman, 1975

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 115
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Oncorhynchus 

kisutch

Coho salmon 6-weeks old 4 R,M,T Mortality 22 – – LC50 1.73 0.38 LC50 calculated by TSK (U.S. 

Environmental Protection Agency, 1991) using nominal concentrations (exposures were reported to have been measured but data were not presented). Control mortalities not reported, assumed to be zero. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-“clean” methods. 

Chapman, 1975

Oncorhynchus 

kisutch

Coho salmon 7-weeks old 4 R,M,T Mortality 22 – – LC50 1.40 .38 LC50 calculated using Probit 

regression (U.S. Environmental Protection Agency, 1992). Exposures reported to be measured but data not presented. Control mortalities not reported, assumed to be zero. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-“clean” methods.

Chapman, 1975

Oncorhynchus 

kisutch

Coho salmon 21-weeks old 4 F,M,T Mortality 24 – – LC50 2.7 .41 Life stage apparently more 

resistant than other life stages. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-“clean” methods.

Chapman, 1975

Oncorhynchus 

kisutch

Coho salmon Adult 4 F,M,T Mortality 24 – – LC50 17.5 .41 Life stage apparently more 

resistant than other life stages. Cu, Pb, and Hg reported above chronic ALC values in dilution water, considered data artifacts due to use of pre-“clean” methods.

Chapman, 1975

Oncorhynchus 

kisutch

Coho salmon Yearling 70 F,M,T Seawater 

adaptation

90 4.5 NOEC 4.5 .55 No effect on capacity to adapt 

to seawater, cortisol levels (a measure of stress) or mortality.

Schreck and Lorz, 

1978

Oncorhynchus 

mykiss

Rainbow trout CSL with fry 100 F,M,T Mortality 217 – – MATC 3.58 .94 Atypical dilution water. Hardness 

amended with MgSO4 only resulting in very low Ca:Mg ratio.

Davies and others, 

1993

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

116 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Oncorhynchus 

mykiss

Rainbow trout CSL with fry 100 F,M,T Mortality 414 – – MATC 3.64 1.40 Atypical dilution water. Hardness 

amended with MgSO4 only resulting in very low Ca:Mg ratio.

Davies and others, 

1993

Oncorhynchus 

mykiss

Rainbow trout CSL with fry 100 F,M,T Mortality 300 8.2 14.2 MATC 10.79 1.15 Atypical dilution water. “Unaged” 

dilution water where highly alkaline water was mixed quickly without allowing equilibrium with Cd.

Davies and 

Brinkman, 1994b

Oncorhynchus 

mykiss

Rainbow trout 0.19 g fry 14 F,M,T Mortality 280 – – ILL 20 1.10 Test at 7.5 ̊C, non-Cd acclimated 

controls.

Stubblefield and 

others, 1999

Oncorhynchus 

mykiss

Rainbow trout Adult 14 F,M,T Mortality 280 – – ILL 6.1 1.10 Test at 7.5 ̊C, non-Cd acclimated 

controls.

Stubblefield and 

others, 1999

Oncorhynchus 

mykiss

Rainbow trout Adult 2 S,M,T Behavior 120 4 – NOEC 4 .65 No change in dominance hierarchies. 

Two treatments (4 and 7 μg/L).

Sloman and others, 

2005

Oncorhynchus 

mykiss

Rainbow trout Adult – S,M,T Behavior 120 – 2 LOEC 2 .65 Decreased competitive ability 

and increased growth in Cd exposed fish. Single exposure concentration.

Sloman and others, 

2003b

Oncorhynchus 

mykiss

Rainbow trout 0.58 g fry 1 S,M,T Behavior 120 3.17 LOEC 3.17 .65 Cd exposed fish were less aggressive 

and subordinate to non-Cd exposed fish.

Sloman and others, 

2003a

Oncorhynchus 

mykiss

Rainbow trout 2.5 g fry 7 F,M,T Behavior 120 .56 2.33 MATC 1.14 .65 Normal antipredator behaviors 

exhibited in response to chemical alarm substance were lost following Cd exposure.

Scott and others, 

2003

Oncorhynchus 

mykiss

Rainbow trout Swimup fry 4 R,U Neurotoxic 163 – 5 LOEC 5 .79 No effect on swimming behavior 

or brain neurochemical activity (cholinesterase (ChE) activity, muscarinic cholinergic receptor (MChR) number, and MChR affinity).

Beauvais and others, 

2001

Oncorhynchus 

mykiss

Rainbow trout Fingerling, 

18 g

4 S,M,T Mortality 50 – – LC50 1.88 .74 4- and 14-day acute values similar. Stubblefield, 1990

Oncorhynchus 

mykiss

Rainbow trout Fingerling, 

18 g

14 F,M,T Mortality 50 – – LC50 1.56 .74 4- and 14-day acute values similar. Stubblefield, 1990

Oncorhynchus 

mykiss

Rainbow trout 5.6 g 30 F,M,T Various 20 – – NOEC .11 .22 No effects on survival, growth, 

swimming, or acclimation.

Hollis and others, 

2000a

Oncorhynchus 

mykiss

Rainbow trout – 30 F,M,T Mortality 25 – – LOEC 2 .25 80 percent mortality in first 5 days. Hollis and others, 

2000b

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 117
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Oncorhynchus 

mykiss

Rainbow trout – 30 F,M,T Physiology 47 – – LOEC 2 0.37 Impaired swimming performance. Hollis and others, 

2000b

Oncorhynchus 

mykiss

Rainbow trout ~1.4 g 30 F,M,T Mortality 140 3 10 MATC 5.48 .72 No effects on survival, growth, 

swimming, or ion composition at NOEC; 30 percent mortality but no other effects at LOEC.

Hollis and others, 1999

Oncorhynchus 

mykiss

Rainbow trout 10-12 g 30 F,M,T Bioaccumulation 140 3 – NOEC 3 .72 No effects on survival, growth; dietary calcium supplementation 

reduced bioaccumulation, but also reduced growth.

Franklin and others, 

2005

Oncorhynchus 

mykiss

Rainbow trout Juvenile 4 F,M,T Mortality 25 – – LC50 2.0 .42 Michael A. Cairns data, as cited 

by Chapman (1982). Chapman (1982) suggested that the data would be suitable for use in criteria derivation and that the complete data may be in Cairns’ thesis. However, his Oregon State University M.S. thesis was located and reviewed but no relevant data were found. Cairns also was located but could provide no additional information. 

Chapman, 1982

Oncorhynchus 

mykiss

Rainbow trout Embryo 

(ELS)

~72 F,M,T Mortality 25 2.6 3.7 MATC 3.10 .25 Michael A. Cairns data, as cited by 

Chapman (1982). See above. 

Chapman, 1982

Oncorhynchus 

mykiss

Rainbow trout Juvenile 120 F,M,T Biochemical 320 1 10 LOEC 10 1.20 Important biochemical functions 

impaired. Exposure concentrations considered too wide to calculate MATC.

Arillo and others, 1984

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

118 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Oncorhynchus 

mykiss

Rainbow trout Adult 540 ? Hatching 102 – 0.2 LOEC 0.2 0.59 Reduced survival (52 vs. 90 percent 

for control) of 4-day old larvae of rainbow trout after parents were exposed to a concentration of 0.2 μg/L at 102 mg/L hardness for 18 months, which is well below the published chronic criterion. The exposed parents had tissue concentrations that were roughly 7 times that of the control fish. This magnitude of effect in an important biological endpoint could justify adjusting the criterion per Stephan and others (1985) recommendations. However, the description of this test was too sparse to answer important questions about the methods and results, such as how the fish were exposed, whether exposure concentrations were measured at sufficient frequency to define the exposure concentrations, and whether treatments were replicated. Because of these uncertainties and because the LOEC is about 10 times lower than any other adverse effect concentrations for this genus, these results probably do not warrant adjusting the chronic criterion.

Birge and others, 1981, 

citing unpublished data by W.J. Birge and J.A. Black

Orconectes 

juvenilis

Crayfish 3rd to 5th 

instars, ~0.02 g

4 R,M,T Mortality 44 – – LC50 60 .67 – Wigginton, 2005

Orconectes 

juvenilis

Crayfish Adult, 4.6 g 4 R,M,T Mortality 44 – – LC50 2,440 .67 Adults were 40 times more resistant 

than newly hatched crayfish.

Wigginton, 2005

Orconectes 

placidus

Placid crayfish 3rd to 5th 

instars, ~0.02 g

4 R,M,T Mortality 55 – – LC50 37 .80 – Wigginton, 2005

Orconectes 

placidus

Placid crayfish Adult, 7 g 4 R,M,T Mortality 44 – – LC50 487 .67 Adults were 13 times more resistant 

than newly hatched crayfish.

Wigginton, 2005

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 119
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Physa integra Snail 6-15 mm 21 F,M,T Mortality 45 3 8 EC10 5 0.36 High control survival and 

concentration-response through day 21, afterwards controls and all treatments suffered high mortalities that did not correspond to concentrations.

Spehar and others, 

1978a

Pimephales 

promelas

Fathead 

minnow

24-48 hours 7 S,M,D Mortality 17 – – LC50 4.4 .20 10- and 14-day exposures more 

sensitive than 7-day exposures. Because longer exposures were available for a given test condition (hardness 17), only the 10- and 14-day results were used in SMAV and hardness regressions.

Suedel and others, 1997

Pimephales 

promelas

Fathead 

minnow

24-48 hours 10 S,M,D Mortality 17 – – LC50 1.6 .20 10- and 14-day exposures more 

sensitive than 7-day exposures. Because longer exposures were available for a given test condition (hardness 17), only the 10- and 14-day results were used in SMAV and hardness regressions.

Suedel and others, 1997

Pimephales 

promelas

Fathead 

minnow

24-48 hours 14 S,M,D Mortality 17 – – LC50 2.3 .20 10- and 14-day exposures more 

sensitive than 7-day exposures. Because longer exposures were available for a given test condition (hardness 17), only the 10- and 14-day results were used in SMAV and hardness regressions.

Suedel and others, 1997

Pimephales 

promelas

Fathead 

minnow

Embryo 

(ELS)

7 S,M,D Mortality 17 4 6 MATC 4.90 .20 10- and 14-day exposures more 

sensitive than 7-day exposures. Because longer exposures were available for a given test condition (hardness 17), only the 10- and 14-day results were used in SMAV and hardness regressions.

Suedel and others, 1997

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

120 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Pimephales 

promelas

Fathead 

minnow

Embryo 

(ELS)

10 S,M,D Mortality 17 1 2 MATC 1.41 0.20 10- and 14-day exposures more 

sensitive than 7-day exposures. Because longer exposures were available for a given test condition (hardness 17), only the 10- and 14-day results were used in SMAV and hardness regressions.

Suedel and others, 1997

Pimephales 

promelas

Fathead 

minnow

Embryo 

(ELS)

14 S,M,D Mortality 17 2 3 MATC 2.45 .20 10- and 14-day exposures more 

sensitive than 7-day exposures. Because longer exposures were available for a given test condition (hardness 17), only the 10- and 14-day results were used in SMAV and hardness regressions.

Suedel and others, 1997

Pimephales 

promelas

Fathead minnowAdult 21 F,M,T Behavior 349 13 18 MATC 15.3 1.26 Cd exposed minnows were more 

vulnerable to predation and displayed altered behavior patterns, including abnormal schooling behavior.

Sullivan and others, 

1978

Potamopyrgus 

antipodarum

New Zealand 

mud snail

Juvenile, 

3-4 mm

2 R,M,T Mortality 148 – 250 LOEC 250 .74 100 percent mortality. Mortality 

was delayed—following exposure for 48 hours and transfer to clean water for observation, after 100 hours in clean water all had died. 48-hour LC50 was 560 μg/L.

Jensen and Forbes, 

2001

Procambarus 

clarkii

Red swamp 

crayfish 

3rd to 5th 

instars, ~0.02 g

4 R,M,T Mortality – – – LC50 42 624 – Wigginton, 2005

Procambarus 

clarkii

Red swamp 

crayfish 

Adult, 18.5 g 4 R,M,T Mortality – – – LC50 53 2,660 Adults were 4 times more resistant 

than newly hatched crayfish.

Wigginton, 2005

Prosopium 

williamsoni

Mountain 

whitefish

Field 

collected, 209 g

14 F,M,T Mortality 50 2.09 4.08 MATC 2.92 .38 Best estimate of a chronic value 

for this genus.

Stubblefield, 1990

Pteronarcys 

dorsata

Stonefly 20-40 mm 28 F,M,T Mortality 45 238 – NOEC 238 .36 – Spehar and others, 

1978a

Ptychocheilus 

lucius

Colorado 

pikeminnow

Larva 4 S,U Mortality 199 – – LC50 78 2.30 No analytical confirmation of 

exposure concentrations.

Buhl, 1997

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 121
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3

)

NOEC 

(μg/L)

LOEC 

(μg/L)

Summary 

statistic

Summary 

effect 

value 

(μg/L)

Corresponding 

acute or 

chronic 

criterion value 

(μg/L)

Notes Reference

Ptychocheilus 

lucius

Colorado 

pikeminnow

Juvenile 4 S,U Mortality 199 – – LC50 108 2.30 No analytical confirmation of 

exposure concentrations.

Buhl, 1997

Rana 

sphenocephala

Southern 

leopord frog

Tadpoles 103 S,M,D Growth and 

mortality

60 – 0.5 LOEC .5 .43 Reduced survival (75 vs. 90 

percent for controls) and a non-significant trend of older and larger metamorphs exposed to Cd through periphyton (about 10 mg/kg dry weight and water (5 to <0.1 μg/L). Adverse effects were noted in the lowest cadmium exposure which was dosed with a single slug resulting in an initial nominal aqueous concentration of 5 μg/L. Aqueous concentrations dropped to <0.1 μg/L during the course of the test. Tissue-residue LOEC was 14.0 mg/kg dry weight (whole body).

James and others, 2005

Salmo trutta Brown trout Swimup fry 4 F,M,T Mortality 29 – – LC50 1.26 .47 4- and 30-day acute values 

similar. LC50s recalculated from author’s data using EPA Probit program.

Brinkman and Hansen, 

2004a

Salmo trutta Brown trout Swimup fry 30 F,M,T Mortality 29 – – LC50 1.08 .47 4- and 30-day acute values 

similar. LC50s recalculated from author’s data using EPA Probit program.

Brinkman and Hansen, 

2004a

Salmo trutta Brown trout Swimup fry 4 F,M,T Mortality 68 – – LC50 3.79 .95 4- and 30-day acute values 

similar. LC50s recalculated from author’s data using EPA Probit program.

Brinkman and Hansen, 

2004a

Salmo trutta Brown trout Swimup fry 30 F,M,T Mortality 68 – – LC50 3.69 0.95 4- and 30-day acute values 

similar. LC50s recalculated from author’s data using EPA Probit program.

Brinkman and Hansen, 

2004a

Salmo trutta Brown trout Swimup fry 4 F,M,T Mortality 151 – – LC50 10.26 1.83 4- and 30-day acute values 

similar. LC50s recalculated from author’s data using EPA Probit program.

Brinkman and Hansen, 

2004a

Salmo trutta Brown trout Swimup fry 30 F,M,T Mortality 151 – – LC50 9.65 1.83 4- and 30-day acute values 

similar. LC50s recalculated from author’s data using EPA Probit program.

Brinkman and Hansen, 

2004a

Salvelinus 

fontinalis

Brook trout Yearling 4 F,M,T Mortality 47 – – LC50 5,080 .71 Yearlings are apparently an 

insensitive life stage.

Holcombe and others, 

1983

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

122 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Species Common name Size or age at test initiation Exposure duration (days) Method Effect Hardness (mg/L as CaCO3) NOEC (μg/L) LOEC (μg/L) Summary statistic Summary effect value (μg/L) Corresponding acute or chronic criterion value (μg/L) Notes Reference

Salvelinus 

fontinalis

Brook trout Fry, 0.21 g 4 S,M,T Mortality 41 – – LC50 0.69 0.63 LC50 estimated as the geometric mean of the lowest exposure 

(0 percent survival at 1.6 μg/L initial measured concentration) and the control treatment (100 percent survival at 0.3 μg/L initial measured concentration). LC50 was not included in table 15 because of data quality uncertainties. Only 2 concentrations were tested (1.5 and 15 μg/L nominal). Cd concentration in control water at 96-hours recorded as at 2.9 μg/L. Authors suggested this may have been an analytical error since 100 percent mortality occurred in the 1.5 μg/L treatment and zero mortaility occurred in controls, however some other treatments also had unexplained increases in Cd during the exposure, raising the possibility of lab contamination. Concentrations in filtered samples were considerably higher than in unfiltered samples, further suggesting sample contamination. 

Carroll and others, 1979

Salvelinus 

fontinalis

Brook trout Sub-adult, 8 

month of

10 R,M,T Mortality 20 8 18 MATC 12.0 .22 LC50 was 14 μg/L. Data point not 

used because older fish may be more resistant than younger fish. Nonstandard duration for either acute or chronic exposures.

Jop and others, 1995

Salvelinus 

namaycush

Lake trout 2-year old 106-112 F,M,T Behavior 81 – .5 LOEC .5 .37 Reduced effectiveness at 

capturing unexposed fathead minnow prey.

Kislalioglu and others, 

1996

Salvelinus 

namaycush

Lake trout 2-year old 240-277 F,M,T Behavior 90 – 0.5 LOEC .5 .55 Reduced effectiveness at 

capturing similarly exposed rainbow trout fingerlings.

Scherer and others, 

1997

Thymallus 

arcticus

Arctic grayling Juvenile 4 S,U Mortality 41 – – LC50 4 .63 No analytical confirmation of 

exposure concentrations.

Buhl and Hamilton, 

1991

Xyrauchen 

texanus

Razorback 

sucker

Larva 4 S,U Mortality 199 – – LC50 139 2.30 No analytical confirmation of 

exposure concentrations.

Buhl, 1997

Xyrauchen 

texanus

Razorback 

sucker

Juvenile 4 S,U Mortality 199 – – LC50 160 2.30 No analytical confirmation of 

exposure concentrations.

Buhl, 1997

Table 17. Other data on effects of cadmium to freshwater organisms.—Continued

Table 17 123
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124 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Appendix – Final Review

Criteria derivation is a complex exercise that should be 

carefully reviewed by rechecking each step of the process and 

the derivation guidelines. Stephan and others (1985, p. 56) 

identified 17 questions that should be especially addressed. 

Since their review questions assumed that the chronic criterion 

would be derived using acute-to-chronic ratios, which was 

not the case here, review questions paralleling those for the 

final acute value were added for the development of the final 

chronic value. The questions and my responses follow:

1. If unpublished data are used, are they well documented?

Several unpublished data sources were used. In each case they 

were available in writing so that the summaries of test values 

and conditions presented here could be reproduced, or the 

data summaries presented here were reviewed and confirmed 

by the original investigators. One unpublished dataset was 

particularly pertinent (matched acute and chronic tests with 

rainbow trout by MA. Cairns). M.A. Cairns was located, but 

no detailed documentation of methods and results could be 

found. These data are presented in table 17 of the main report 

(Other data) but not directly used in criteria calculations.

2. Are all required data available?

Yes, the 8-family minimum species diversity rules were met 

for both the acute and chronic datasets.

3. Is the range of acute values for any species greater than a 

factor of 10?

Yes. Daphnia magna acute values varied by more than a factor 

of 100, even after hardness-adjustment. However, except for 

clearly resistant genotypes, this variability was considered 

representative of the variability found in field populations. 

Whether this greater range of values for Daphnia results is 

an artifact of its larger dataset or is unique to Daphnia and 

cadmium is not known. Available Daphnia magna values with 

copper suggest less variability than with cadmium. Biotic 

ligand model adjusted acute values varied by a maximum 

factor of 30, with most values varying by less than a factor of 

10 (U.S. Environmental Protection Agency, 2003a)

4. Is the range of Species Mean Acute Values for any 

genus greater than a factor of 10?

No, the greatest range (Daphnia) was a factor of 5.

5. Is there more than a factor of 10 difference between the 

four lowest Genus Mean Acute Values?

No, differences were less than a factor of 2.

6. Are any of the four lowest Genus Mean Acute Values 

questionable?

Not in the final iteration. Earlier, one of the lowest GMAVs 

(Striped bass, Morone saxatilis, now listed in table 17) 

was based on unmeasured exposures. Since this seemed 

questionable, and no obvious basis for selectively removing 

values reported from unmeasured exposures, all test values 

based upon unmeasured exposures were deleted from the acute 

and chronic datasets.

7. Is the Final Acute Value reasonable in comparison with 

the Species Mean Acute Values and Genus Mean Acute 

Values?

Yes, when plotted as a species-sensitivity distribution and fit to 

a logistic curve (not shown), the data were evenly distributed. 

The 5th percentile FAV (2.41 μg/L) fell between the GMAVs 

with cumulative probabilities of the 4th and 6th percentiles (2.1 

and 2.6 μg/L respectively).

8. For any “important” or “critical” species3, is the 

geometric mean of the acute values from tests in which the 

concentrations of test material were measured lower than 

the Final Acute Value4?

Yes, species mean acute values for cutthroat trout, rainbow 

trout, and bull trout (all based on measured values) were all 

lower than the calculated FAV. The FAV was lowered to the 

lowest of these SMAVs.

9. Are any of the chronic values questionable?

Of the values retained for use, none were obviously suspect. 

Although few if any of the chronic test values could be 

considered ideal and the results often raised questions, none 

of the table 16 values seemed so “questionable’ as to be 

discarded. Assigning values to tests with unbounded low 

effects were particularly troublesome (i.e., adverse effects 

were observed at the lowest exposure). Taking a geometric 

mean of the low-effect concentration and the control was 

usually unsatisfactory because the control was often a “less 

than” value too. In these situations, chronic values were 

selected on a case-by-case basis considering the magnitude of 

effect and other values for the species if any.

3

 As described in methods under “Extrapolating small toxicity test datasets 

to aquatic communities,” Stephan and others (1985) limited their consideration 

of “important” species in this context to commercially or recreationally 

important species. However, in the context of site-specific criteria, Stephan 

and others (1994) defined the term “critical” species more broadly and that 

broader definition seems more appropriate here. 

4

 Stephan and others (1985, p. 56) limited this question to “acute values 

from flow-through tests in which concentrations were measured” however 

since renewal and static tests in which concentrations were measured were 

also used to develop the Final Acute Value this question was modified (see the 

section “Flow through, renewal, and static test exposures”).

Appendix 125
10. Is the range of chronic values for any species greater than 

a factor of 10?

As with acute data, Daphnia magna values were highly 

variable. Chronic values varied by over a factor of 20. As with 

the acute data, this variability is assumed to at least partly 

reflect that found in field populations. Whether this high 

variability in the sensitivity of Daphnia magna to cadmium 

reflects inherent differences from that of other species or 

chemicals or if the difference is an artifact of the intensive 

studies with Daphnia magna and cadmium is unknown. In 

EPA’s (2003a) compilation of copper results, chronic values 

for Daphnia magna only varied by about a factor of three. No 

values were excluded solely due to differences from other 

values.

11. Is the range of Species Mean Chronic Values for any 

genus greater than a factor of 10?

Yes, within the genus Salvelinus the SMCV for lake trout 

was 10.3X higher than the SMCV for bull trout. Because 

the difference was greater than a factor of 10, the lake trout 

SMCV was based upon a single test, and because the source 

of that test for lake trout data included no information on what 

percent reductions in survival corresponded to the NOEC 

or LOEC values, the lake trout value was excluded from the 

GMCV calculation. Otherwise the greatest range was a factor 

of 6.6 for Daphnia.

12. Is there more than a factor of 10 difference between the 

four lowest Genus Mean Chronic Values?

No, there was only about a factor of 3.6 difference between the 

lowest and 4th lowest GMCVs. .

13. Are any of the four lowest Genus Mean Chronic Values 

questionable?

As with question number 6, not in the final iteration. Earlier, 

one of the lowest GMCVs (cladoceran, Moina macrocopa, 

was based on unmeasured exposures (now listed in table 17, 

for “Other data.” 

14. Are chronic values available for acutely sensitive 

species?

Yes. Chronic values are available for species within each of 

each of the four most acutely sensitive genera. No chronic 

values are available for the most acutely sensitive species, 

cutthroat trout Oncorhynchus clarki. However, chronic data 

are available for two other species of Oncorhynchus and are 

assumed to be reasonably representative of chronic sensitivity 

for other members of the genus Oncorhynchus (at least more 

so than values for species from other genera). Cutthroat trout 

and rainbow trout SMAVs were similar, with overlapping 

acute values.

15. Is the range of acute-chronic ratios greater than a factor 

of 10?

Yes. The ACRs ranged from 1.1 to 424. Since the species 

mean acute-chronic ratios tended to increase with species 

mean acute values, acute-chronic ratios were calculated for 

species with species mean acute values that were close to 

the final acute value. To judge how close “close” should be, 

species mean acute values that were within ~10X of the final 

acute value were used. This resulted in a final ACR that was 

similar to the FAV–FCV ratio. Although presented in the 

report, acute-chronic ratios were not used in derivation of the 

chronic cadmium criterion but are presented in the report.

16. Is the Final Chronic Value reasonable in comparison with 

the available acute and chronic data?

Yes. The final chronic value fell slightly above the most 

sensitive species with a cumulative probability value of 0.05. 

The chronic species sensitivity distribution was fairly even and 

fit well to a curve (fig. 5). A few acute values from the most 

sensitive genus Oncorhynchus are close to the final chronic 

value and one was lower. However, these were the extremes of 

the distribution of available values for this genus. Overall, the 

most sensitive species mean acute value (cutthroat trout) was 

3.4X greater than the final chronic value, which is probably 

enough higher that minimal acute toxicity would result at FCV 

concentrations. 

17. Is the measured or predicted chronic value for any 

“critical” species below the Final Chronic Value?

No, unless Hyalella azteca is considered “critical.” 

18. Are any of the “other data” important?

Possibly. Two independent tests found that the ability of 

Salvelinus trout to capture prey was impaired at exposures at 

or below chronic criterion concentrations. However, a longterm whole lake ecosystem experiment studied cadmium 

exposed Salvelinus populations without finding any evidence 

of reduced body conditions, making the extrapolation of these 

tests to the wild uncertain. Regardless, these “other data” 

caution that careful bioassessment of fish health, populations 

and fish and macroinvertebrate assemblages would be prudent 

in situations where cadmium concentrations in receiving 

waters approach criterion concentrations.

126 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
19. Do any data look like they might be outliers?

Not in the final compilation of acute and chronic values that 

seemed acceptable for use (tables 15 and 16). Several values 

appeared to be outliers and thus were given extra scrutiny. 

Based on this scrutiny, values from several tests were not used 

in the FAV or FCV calculations because they appeared to have 

tested life stages that were not the most sensitive, organisms 

were pre-exposed to cadmium, or otherwise questionable data 

(e.g., Chironomus riparius, brook trout, or coho salmon, table 

17). One value for Ceriodaphnia dubia was not used solely 

because it was considered an outlier (table 16). 

20. Are there any deviations from the guidelines? Are they 

acceptable?

There were three deviations from the Guidelines. (1) The 

guidelines gave flow-through toxicity tests priority over 

static tests. Here, flow-through, renewal, and static test 

results were all used with equal priority assuming that they 

are otherwise acceptable (for example, used sensitive life 

stages, provided sensitive results in comparison to other 

tests within the genus, and concentrations were measured). 

(2) The guidelines would accept results of tests in which 

the exposure concentrations were not measured if no better 

data for a species were available. Here no test results based 

upon unmeasured concentrations were used since sufficient 

values with measured concentrations to meet minimum 

diversity requirements were available. (3) The guidelines 

gave preference to results of chronic tests conducted over an 

organism’s full life-cycle over shorter-term tests. Here shorterterm tests also were used if results were more sensitive, or 

similar in sensitivity to results from life-cycle tests. See “Data 

acceptability review.” 

These deviations from the guidelines were each supported by 

a technical rationale. Considering the data available in 2005, 

but not available in 1984 when the guidelines were written, the 

deviations seemed more appropriate than did rigid adherence 

to the guidelines. Still, these deviations from the guidance 

could be debated; Three peer reviewers commented that they 

concurred with the deviations; one commented that they did 

not fully agree with the deviations.

21. On the basis of all available pertinent laboratory and 

field information, are the criteria consistent with sound 

scientific evidence? If they are not, other criteria, either 

higher or lower should be derived using appropriate 

modifications to criteria derivation Guidelines.

Most data indicate that measurable, adverse biological effects 

are unlikely in lake or stream ecosystems that only rarely 

exceed the criteria values calculated here. Behavioral data 

with Salvelinus (lake and brook trout) are ambiguous, with 

adverse effects demonstrated in laboratory settings but with 

no obvious adverse effects observed in a long-term whole lake 

experiment. Because of this uncertainty, careful biomonitoring 

and assessment of fish health, and benthic macroinvertebrate 

or zooplankton assemblages in streams or lakes, respectively, 

could be important if discharges were to cause receiving 

waters to approach criteria conditions.

Appendix 127
Glossary

Acute toxicity Acute toxicity tests are generally used to 

determine the concentration of test material that produces 

a specific adverse effect on a specified percentage of test 

organisms during a short exposure. Because death is an 

obviously important adverse effect and is easily detected for 

many species, the most common acute toxicity test is the 

acute lethality test. Experimentally, effects on 50 percent 

of a group of test organisms are the most reproducible 

and easily determined measure of toxicity, and 96 hours is 

often a convenient, useful exposure duration. Therefore, 

the measure of acute toxicity most often used with fishes, 

macroinvertebrates, and amphibians is the 96-hour LC50. 

However, because immobilization is a severe effect and is not 

easy to distinguish from death for some species, the measure 

of acute toxicity most often used with daphnids and midge 

larvae is the 48-hour EC50 based on death plus immobilization 

(American Society for Testing and Materials, 1997). 

ACR Acute to chronic ratio, used for estimating chronic 

effect values from acute effects data.

Alevin Newly hatched fish that still contain a yolk sac. 

Summary of life stage terms: alevins are newly hatched fish 

that still contain a yolk sac. Swim-up fry are fish that are 4–6 

weeks post hatching, have resorbed their yolk sacs and have 

begun feeding, and have just swum up from their incubation 

cups or natural substrate to swim swimming freely in the water 

column. For the salmonids, larger juveniles that may be up to 

two years old are often referred to as parr. Juvenile salmon and 

anadromous trout that are physiologically adapted to seawater 

are called smolts.

BAF Bioaccumulation factor. A BAF is intended to 

account for net uptake from both food and water in a realworld situation. A BAF almost has to be measured in a field 

situation under reasonably steady-state conditions in which 

predators accumulate the material directly from the water and 

by consuming prey that itself could have accumulated the 

material from both food and water.

BCF Bioconcentration factor. A BCF is the quotient of the 

concentration of a material in one or more tissues of an aquatic 

organism divided by the average concentration in the solution 

in which the organism had been living. A BCF is intended 

to account only for net uptake directly from water and thus 

almost has to determined through laboratory testing.

CCC Criterion continuous concentration, synonymous with 

“chronic criterion.”

Chronic toxicity As used here, chronic toxicity generally 

refers to adverse effects to organisms during a long-term 

exposure to concentrations of test material. Effects that are 

considered “adverse” include lethal or sublethal effects that 

can be related to survival and reproduction rates. The duration 

of “long-term” exposures varies by species and test protocol, 

but must be longer than that used for acute exposures and long 

enough to measure sublethal effects such as growth. “Longterm” exposures may be as brief as seven days in short-term 

tests for estimating chronic toxicity. Unlike “acute toxicity,” 

the usage of the term “chronic toxicity usage varies in the 

ecotoxicology literature. See also “Life-cycle versus shorterterm “chronic” data” and “Statistical interpretation of chronic 

tests.”

CMC Criterion maximum concentration, synonymous with 

“acute criterion.” 

Critical species In the context of deriving site specific 

water-quality criteria and related assessments that use 

the SSD concept, “critical species” have been defined to 

include 1) keystone species that are of great ecological value 

because their loss would indirectly affect many other species 

or ecosystem function; 2) species of special conservation 

status through treaties, laws, or policies as threatened, 

endangered, or vulnerable; 3) recreationally valued species; 

or 4) commercially important species (U.S. Environmental 

Protection Agency, 1994; Posthuma and others, 2002; Stephan 

and others, 1994b).

Chronic limits “A chronic value may be obtained by 

calculating the geometric mean of the lower and upper chronic 

limits from a chronic test or by analyzing chronic data using 

regression analysis. A lower chronic limit is the highest tested 

concentration (a) in an acceptable chronic test, (b) which 

did not cause an unacceptable amount of adverse effect on 

any specified biological measurements, and (c) below which 

no tested concentration caused an unacceptable effect. An 

upper chronic limit is the lowest tested concentration (a) in an 

acceptable chronic test, (b) which did cause an unacceptable 

amount of adverse effect on any specified biological 

measurements, and (c) above which all tested concentration 

also caused such an effect.” (Stephan and others 1985, p. 39)

Dissolved metals Operationally defined as those metals 

that pass through a 0.45 μm filter, which includes metals 

chemically dissolved in water and very small colloidal 

particles. 

DOM Dissolved organic matter.

dw Dry weight.

ECp Concentration causing effects in p percentage of the 

test population, such as the EC50, EC20, EC10, and so on.

ELS Early-life stage toxicity test. Embryos and subsequent 

larvae of fish species are maintained for 28 to 120 days, 

depending on species in treatments of water with differing 

concentrations of test materials added. 

EPA U.S. Environmental Protection Agency

FAV Final acute value

FCV Final chronic value 

128 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Flow-through toxicity test A test in which the test solution 

flows through the test chamber on a once-through basis 

throughout the test. Compare with renewal or static toxicity 

tests. Usage note: The term “flow-through” could imply a 

test with flowing water that is constructed to mimic a stream 

environment. This is not the case; rather the term refers to 

the once-through, continuous delivery of test solutions (or 

frequent delivery in designs using a metering system that 

cycles every few minutes). Flows on the order of about 

5-volume replacements per 24 hours are insufficient to 

cause discernable flow velocities. In this report, tests that are 

designed to mimic stream environments are called “artificial 

stream” tests.

Fry Juvenile fish that have resorbed their yolk sac. See 

listing of fish life stage terms under “alevin.”

GMAV Genus mean acute value, calculated as the geometric 

mean of all SMAVs for that genus.

GMCV Genus mean chronic value, calculated as the 

geometric mean of all SMCVs for that genus.

ILL Incipient lethal level. The concentration of a chemical 

that is lethal to 50 percent of the test organisms as a result 

of exposure for periods sufficiently long that acute lethal 

action has essentially ceased. Also may be referred to as a 

time-independent LC50, incipient LC50, or asymptotic LC50 

concentration.

IBI Index of biotic integrity. IBIs are empirical additive 

indexes consisting of various metrics describing an assemblage 

such species composition, diversity, functional organization, 

presence or absence of pollution sensitive or tolerant taxa 

(Karr, 1991).

IDEQ Idaho Department of Environmental Quality.

Instar A stage of an insect or other arthropod between 

molts.

JGS Juvenile growth and survival test; a chronic test 

initiated with fish as swim-up fry or with invertebrates such as 

amphipods shortly after hatching.

LC Life cycle test 

LC50 Concentration lethal to 50 percent of the organisms in 

a toxicity test.

LOEC Lowest-observed-effect concentration, see NOEC.

MATC A maximum acceptable toxicant concentration 

(MATC) is a hypothetical test concentration which is assumed 

to be a threshold for toxic effects. Its point estimate is the 

geometric mean of the NOEC and the LOEC. 

NAWQA USGS National Water Quality Assessment 

Program

NOEC As used in this report, the no-observed-effect 

concentration (NOEC) from a toxicity test is the highest test 

concentration that results in responses that are not appreciably 

different from the control responses. The lowest-observedeffect concentration (LOEC) is the first treatment that is 

greater than the NOEC. Note: this usage is slightly different 

from that often used elsewhere where the NOEC is often 

defined as the concentration that is not statistically different 

from controls. See text section on text “Interpretation of 

chronic effects” 

Parr Juvenile salmonids that are freshwater adapted and 

usually between 2-months and 2-years in age. See listing of 

different fish life stages under “alevin.”

Renewal toxicity test The “renewal” technique is like 

the static technique except that test organisms are periodically 

exposed to fresh test solution of the same composition, usually 

once every 24 h or 48 h, by replacing nearly all the test 

solution.

SMAV Species mean acute value, calculated as the 

geometric mean of all acceptable acute values for that species, 

following the adjustment of each acute value to a common 

hardness value using observed hardness-toxicity relations.

SMCV Species mean chronic value, calculated in the same 

way as SMAVs.

SSD Species sensitivity distribution, statistical distribution 

estimated from toxicity data for a dataset of species responses 

and are considered to be a sample of an entire population (in 

the statistical sense) of species responses. Note: SSDs should 

more precisely be called “taxa sensitivity distributions” since 

they may be developed at the species, genus, family or other 

taxonomic level. However, the common usage seems to be 

“SSD” regardless of taxonomic level, and that usage is kept 

in this report.Static toxicity test In “static” exposures, test 

solutions and organisms are placed in chambers and kept there 

for the duration of the test. 

Swim-up fry Fry that have left their nests in or on natural 

substrates or their artificial incubation cups and are freely 

swimming in the water column and feeding. See listing of 

different fish life stages under “alevin.”

Total concentration Metals concentration determined from 

an unfiltered sample. 

TRV Tissue reference value. A TRV is considered as a tissue 

chemical concentration above which adverse effects might 

occur to aquatic species following chronic exposure and below 

which it is unlikely that adverse effects will occur.

USGS U.S. Geological Survey 

Glossary 129
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130 Cadmium Risks to Freshwater Life: Derivation and Validation of Low-Effect Criteria Values 
Manuscript approved for publication, September 21, 2006

Prepared by the USGS Publishing Network,

Publishing Service Center, Tacoma, Washington

Bill Gibbs

Bob Crist

Debra Grillo

Sharon Wahlstrom

Bobbie Jo Richey

For more information concerning the research in this report, contact the

Director, Idaho Water Science Center,

U.S. Geological Survey, 230 Collins Road

Boise, Idaho 83702

http://id.water.usgs.gov
SIR 2006–5245, Version 1.2 Mebane

Cadmium Risks to Freshwater Life: Derivation and Validation of 

Low-Effect Criteria Values using Laboratory and Field Studies