oEPA
United States
Environmental Protection
Agency
Office of Water
4304T
EPA 822-R-21-006
August 2021
2021 Revision* to:
Aquatic Life Ambient Water
Quality Criterion for
Selenium - Freshwater
2016
* This 2021 revision entails errata regarding footnotes on pages xv and 99, as
described on page vii.
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, D.C.

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Table of Contents
Table of Contents	ii
List of Tables	v
List of Figures	vi
Notice	vii
Foreword	viii
Acknowledgements	ix
Acronyms	xi
Executive Summary	xii
1	Introduction and Background	1
1.1 History of the EPA Recommended Selenium AWQC for Aquatic Life	1
2	Problem Formulation	4
2.1	Overview of Selenium Sources and Occurrence	4
2.2	Environmental Fate and Transport of Selenium in the Aquatic Environment	8
2.2.1	Selenium Species in Aquatic Systems	8
2.2.2	Bioaccumulation of Selenium in Aquatic Systems	10
2.3	Mode of Action and Toxicity of Selenium	12
2.4	Narrow Margin between Sufficiency and Toxicity of Selenium	14
2.5	Interactions with Mercury	15
2.6	Assessment Endpoints	16
2.7	Measures of Effect	19
2.7.1	Fish Tissue	20
2.7.2	Water	23
2.7.3	Summary of Assessment Endpoints and Measures of Effect	24
2.7.4	Conceptual Model of Selenium Effects on Aquatic Life	25
2.7.5	Analysis Plan for Derivation of the Chronic Fish Tissue-Based Criterion
Elements	26
2.7.6	Analysis Plan for Derivation of the Fish Tissue Criterion Elements Duration	27
2.7.7	Analysis Plan for Derivation of the Fish Tissue Criterion Elements Return
Frequency	28
2.7.8	Analysis Plan for Derivation of Chronic Water-based Criterion Element	29
2.7.9	Analysis Plan for Derivation of the Water Criterion Elements Duration	31
2.7.10	Analysis Plan for Intermittent-Exposure Water-based Criterion Element
Derivation	32
3	Effects Analysis for Freshwater Aquatic Organisms	34
3.1 Chronic Tissue-Based Selenium Criterion Element Concentration	34
3.1.1	Acceptable Studies of Fish Reproductive Effects for the Four Most Sensitive
Genera	35
3.1.2	Summary of Acceptable Studies of Fish Reproductive Effects	44
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3.1.3	Derivation of Tissue Criterion Element Concentrations	47
3.1.4	Invertebrate Chronic Effects	54
3.1.5	Summary of Relevant Invertebrate Tests	57
3.1.6	Selenium Fish Tissue Toxicity Data Fulfilling Minimum Data Needs	58
3.2	Chronic Water Column-based Selenium Criterion Element	60
3.2.1	Translation from Fish Tissue Concentration to Water Column Concentration	60
3.2.2	Equation Parameters	69
3.2.3	Food-Web Models	81
3.2.4	Classifying Categories of Aquatic Systems	82
3.2.5	Deriving Protective Water Column Concentrations for Lentic and Lotic System
Categories	87
3.2.6	Derivation of Averaging Period for Chronic Water Criterion Element	93
3.3	Intermittent-Exposure Water Criterion Element: Derivation from the Chronic Water
Criterion Element	94
4	National Criterion for Selenium in Fresh Waters	98
4.1 Protection of Downstream Waters	102
5	Site-specific Criteria	103
6	Effects Characterization	104
6.1	Fish and Amphibians	104
6.1.1	Principles for Using Studies for which ECios Cannot Be Calculated	104
6.1.2	Acceptable Studies of Fish Reproductive Effects of Genera that were not among
the Four Most Sensitive Genera	105
6.1.3	Reproductive Effects in Catfish (Ictaluridae)	Ill
6.1.4	Reproductive Effects in Amphibians (Xenopus laevis)	116
6.1.5	Reproductive Studies Not Used in the Numeric Criterion Derivation	116
6.1.6	Salmo GMCV: EPA Re-analysis of a Key Study Used in Criterion Derivation	119
6.1.7	Impact of Number of Tested Species on Criterion Derivation	120
6.1.8	Comparisons between Concentrations in Different Tissues	120
6.1.9	Studies of Non-Reproductive Effects	121
6.1.10	Special conditions for consideration of primacy of water column criterion
elements over fish tissue criterion elements	126
6.2	Water	127
6.2.1	Validation of Translation Equation for Developing Water Column
Concentrations	127
6.2.2	Sulfate-Selenium Interactions	130
6.3	Uncertainty	131
6.3.1	Tissue Criterion Element	131
6.3.2	Trophic Transfer Factors	136
6.3.3	Enrichment Factors	138
6.3.4	Water Values	140
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6.4	Protection of Threatened or Endangered Species	140
6.4.1 Special Consideration for Pacific Salmonid Juveniles	142
6.5	Aquatic-Dependent Wildlife is Beyond the Scope of this Aquatic Criteria
Derivation	145
6.6	Summary	146
7 References	149
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List of Tables
Table 1. Summary of the Recommended Freshwater Selenium Ambient Chronic Water Quality
Criterion for Protection of Aquatic Life	xv
Table 2.1. Predominant Chemical Forms of Selenium in Discharges Associated with Different
Activities and Industries	8
Table 2.2. Summary of Assessment Endpoints and Measures of Effect Used in Criteria
Derivation for Selenium	24
Table 3.1. Maternal Transfer Reproductive Toxicity Studies	45
Table 3.2. Ranked Genus Mean Chronic Values for Fish Reproductive Effects Measured as
Egg or Ovary Concentrations	47
Table 3.3. Four Lowest Genus Mean Chronic Values for Fish Reproductive Effects	48
Table 3.4. Tested Reproductive-Effect Whole Body (WB) Concentrations Measured Directly
or Converted to WB Concentrations from Egg-Ovary (EO) Concentrations	50
Table 3.5. The Lowest Four Reproductive-Effect Whole-Body GMCVs	51
Table 3.6. Tested Reproductive-Effect Muscle (M) Concentrations Measured Directly or
Converted to M Concentrations from Egg-Ovary (EO) Concentrations	52
Table 3.7. The Lowest Four Reproductive-Effect Fish Muscle GMCVs	53
Table 3.8. Ranked Invertebrate Whole-Body Chronic Values with Translation to Expected
Accompanying Fish Whole-Body Concentrations	58
Table 3.9. Minimum Data Requirements Summary Table Reflecting the Number of Species and
Genus Level Mean Values Represented in the Chronic Toxicity Dataset for Selenium in
Freshwater	58
Table 3.10. EPA-Derived Trophic Transfer Factor (TIT) Values for Freshwater Aquatic
Invertebrates	75
Table 3.11. EPA-Derived Trophic Transfer Factor (TTF) Values for Freshwater Fish	76
Table 3.12. EPA-Derived Egg-Ovary to Whole-Body Conversion Factor (CF) Values	79
Table 3.13. Data for the 65 Site Minimum Translations of the Egg-Ovary Criterion
Concentration Element to a Water Column Concentration	89
Table 3.14. Water Column Criterion Element Concentration Values Translated from the Egg-
Ovary Criterion Element	92
Table 3.15. Representative Values of the Intermittent Water Criterion Concentration
Element	97
Table 4.1. Summary of the Recommended Freshwater Selenium Ambient Chronic Water
Quality Criterion for Protection of Aquatic Life	99
Table 6.1. Correlation Matrix (Values of r) for Ictaluridae and Centrarchidae Abundance and
for Selenium Food Chain Contamination for the Hyco Reservoir	112
Table 6.2. Freshwater Chronic Values from Acceptable Tests - Non-Reproductive
Endpoints	122
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List of Figures
Figure 2.1. Selenium in Surficial Soils and Aquatic Sediments in counties of the Conterminous
United States	6
Figure 2.2. Diagram of Selenium Partitioning, Bioaccumulation, and Effects in the Aquatic
Environment	25
Figure 2.3. Conceptual Model for Translating the Selenium Egg-Ovary Concentration to a
Water Column Concentration	32
Figure 3.1. Distribution of Reproductive-Effect GMCVs for Fish Measured as Egg-Ovary
Concentrations	49
Figure 3.2. Distribution of Reproductive-Effect GMCVs for Fish, either Measured as Whole-
Body Concentrations in the Original Tests, or Measured as Egg-Ovary Concentrations
but Converted to Whole-Body	51
Figure 3.3. Distribution of Reproductive-Effect GMCVs for Fish, either Measured as Muscle in
the Original tests, or Measured as Egg-Ovary Concentrations but Converted to Muscle
Concentrations	54
Figure 3.4. Example Aquatic System Scenarios and the Derivation of the Equation Parameter
TTp	gy
Figure 3.5. Effect of Relative Sample Collection Time on Correlation Coefficients of Selenium
Measurements in Particulate Material, and Invertebrate and Fish Tissue	72
Figure 3.6. Enrichment Factors (EF) for 96 Aquatic Sites Derived from Published Studies and
Organized by Waterbody Type	84
Figure 3.7. The Relationship between Cwater and Cparticuiate, and Cwater and EF for the 39 Lentic
and 57 Lotic Aquatic Systems	86
Figure 3.8. Distribution of EF Values for the Same 96 Aquatic Systems	87
Figure 3.9. Probability Distribution of the Water Column Concentrations Translated from the
Egg-Ovary Criterion Element at 26 Lentic and 39 Lotic Aquatic Sites	92
Figure 3.10. Illustration of Intermittent Spike Exposure Occurring for a Certain Percentage of
Time (e.g., 10%) Over a 30-Day Period, and Background Exposure Occurring for the
Remaining Percentage of Time (e.g., 90%)	95
Figure 6.1. Crutchfield (2000) Observations of Channel Catfish (CCF) and Largemouth Bass
(LMB) in Hyco Reservoir Beginning a Few Years after Populations of Largemouth
Bass had been Reduced by Se Contamination	115
Figure 6.2. Distribution of Fish Reproductive Effect GMCVs from Figure 3.2 and
Distribution of Fish Nonreproductive Effect GMCVs and Invertebrate GMCVs	125
Figure 6.3. Scatter Plot of Predicted Versus Measured Concentrations of Selenium in Fish	129
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ERRATA
In 2021, EPA identified that the following text was missing from the second sentence in
footnote 4 in the criterion table: "When selenium inputs are increasing," Corrected footnote 4
now states: "4. Water column values are based on dissolved total selenium in water and are
derived from fish tissue values via bioaccumulation modeling. When selenium inputs are
increasing, water column values are the applicable criterion element in the absence of steady-
state condition fish tissue data." Footnotes 2 and 3 also reflect that footnote 4 was corrected.
Notice
This document has undergone a contractor-led external expert peer-review, as well as an
EPA review process following publication and public comments received on the May 14, 2014,
and July 28, 2015 criteria drafts. Final review by the Health and Ecological Criteria Division,
Office of Science and Technology, U.S. Environmental Protection Agency, has been completed,
and the document has been approved for publication.
This document provides guidance to States and Tribes authorized to adopt water quality
standards under the Clean Water Act (CWA), to protect aquatic life from toxic effects of
selenium. Under the CWA, States and Tribes are to adopt water quality criteria to protect
designated uses. State and tribal decision makers retain the discretion to adopt approaches on a
case-by-case basis that differ from this guidance when appropriate. While this document contains
EPA's scientific recommendations regarding ambient concentrations of selenium that protect
aquatic life, it does not substitute for the CWA or EPA's regulations; nor is it a regulation itself.
Thus, it cannot impose legally-binding requirements on EPA, States, Tribes, or the regulated
community, and might not apply to a particular situation based upon the circumstances. EPA
may change this document in the future. This document has been approved for publication by the
Office of Science and Technology, Office of Water, U.S. Environmental Protection Agency.
Mention of trade names or commercial products does not constitute endorsement or
recommendation for use. This document can be downloaded from:
http://www.epa.gov/waterscience/criteria/aqlife.html
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Foreword
Section 304(a)(1) of the Clean Water Act requires the Administrator of the Environmental
Protection Agency to publish water quality criteria that accurately reflect the latest scientific
knowledge on the kind and extent of all identifiable effects on health and welfare that might be
expected from the presence of pollutants in any body of water, including ground water. This
document presents EPA's updated chronic ambient water quality criterion (AWQC) for the
protection of aquatic life based upon consideration of all available information relating to effects
of selenium on aquatic organisms. EPA has incorporated revisions into this final document based
on comments from the general public and an external expert peer review panel on an earlier draft
published in the Federal Register in May 14, 2014, and comments from the general public on a
second draft published in the Federal Register in July 28, 2015.
The term "water quality criteria" is used in two sections of the Clean Water Act, section
304(a)(1) and section 303(c)(2). The term has a different program impact in each section. In
section 304, the term represents a non-regulatory, scientific assessment of ecological effects. The
criterion presented in this document is such a scientific assessment. If water quality criteria
associated with specific designated uses are adopted by a state or authorized tribe as water
quality standards under section 303, and approved by EPA, they become applicable Clean Water
Act water quality standards in ambient waters within that state or tribe. Water quality criteria
adopted in state or tribal water quality standards could have the same numerical values as criteria
developed under section 304. However, states and authorized tribes may adopt water quality
criteria that reflect adjustments to EPA's recommended section 304 criteria to reflect local
environmental conditions and human exposure patterns. Alternatively, states and authorized
tribes may derive numeric criteria based on other scientifically defensible methods but the
criteria must be protective of designated uses. It is not until their adoption as part of state or
tribal water quality standards, and subsequent approval by EPA, that criteria become Clean
Water Act applicable water quality standards. Guidelines to assist the states and authorized tribes
in modifying the criteria presented in this document are contained in the Water Quality Standards
Handbook (U.S. EPA 1994, as updated), which along with additional guidance on the
development of water quality standards and other water-related programs of this Agency have
been developed by the Office of Water.
This document provides guidance only. It does not establish or affect legal rights or
obligations. It does not establish a binding norm and cannot be finally determinative of the issues
addressed. Agency decisions in any particular situation will be made by applying the Clean
Water Act and EPA regulations on the basis of specific facts presented and scientific information
then available.
Elizabeth Southerland
Director
Office of Science and Technology
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Acknowledgements
Technical Analysis Leads
Joseph Beaman, U.S. EPA, Office of Water, Office of Science and Technology, Health and
Ecological Criteria Division, Washington, DC
beaman. i oe@epa. gov (primary contact person)
Gary Russo, U.S. EPA, Office of Water, Office of Science and Technology, Standards and
Health Protection Division, Washington, DC
russo.gary@epa.gov
Charles Delos (Retired) U.S. EPA, Office of Water, Office of Science and Technology, Health
and Ecological Criteria Division, Washington, DC
U.S. EPA Office of Water Reviewers
Elizabeth Behl
Kathryn Gallagher
Lisa Huff
Michael Elias
Intra-Agency Panel Peer Reviewers (2014-2016)
Russell Erickson, Dale Hoff and Charles Stephan, U.S. EPA, Office of Research and
Development, Mid-Continent Ecology Division, Duluth, MN
Cindy Roberts, U.S. EPA, Office of Research and Development, Office of Science Policy,
Washington, DC
Jim Lazorchak, U.S. EPA, Office of Research and Development, National Exposure Research
Laboratory, Cincinnati, OH
Jeff Gallagher, U.S. EPA, Office of Chemical Safety and Pollution Prevention, Office of
Pollution Prevention and Toxics, Washington, DC
Catherine Aubee and Geoff Sinclair, Office of Chemical Safety and Pollution Prevention, Office
of Pesticide Programs, Arlington, VA
David Hair, Laura Phillips, and Scott Wilson, U.S. EPA, Office of Water, Office of Wastewater
Management, Washington, DC
Rosaura Conde, Ruth Chemerys, and Eric Monschein, U.S. EPA, Office of Water, Office of
Wetlands, Oceans, and Watersheds, Washington, DC
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Jim Keating. Julianne McLaughlin, and Lars Wilcut and, U.S. EPA, Office of Water, Office of
Science and Technology, Washington, DC
Cheryl Atkinson and Frank Borsuk, U.S. EPA Region 3, Philadelphia, PA, and Wheeling, WV
Joel Hansel, U.S. EPA Region 4, Atlanta, GA
Dean (Robie) Anson, Candice Bauer, Katharine Marko, and Angela Vincent, U.S. EPA Region
5, Chicago, IL
Jason Gildea and Lareina Guenzel, U.S. EPA Region 8, Denver, CO
Diane Fleck, Eugenia McNaughton, and Daniel Oros, U.S. EPA Region 9, San Francisco, CA
Lisa Macchio, and Burt Shepard, U.S. EPA Region 10, Seattle, WA
Technical support was provided by Great Lakes Environmental Center
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Acronyms
AE	Assimilation Efficiency
AWQC	Ambient Water Quality Criteria
BAF	Bioaccumulation Factor
CF	Conversion Factor
CV	Chronic Value (expressed in this document as an EC 10)
CWA	Clean Water Act
dw	Dry Weight
ECx	Effect Concentration at X Percent Effect Level
EF	Enrichment Factor
EPA	Environmental Protection Agency
EO	Egg Ovary
FCV	Final Chronic Value
GMCV	Genus Mean Chronic Value
HNOEC	Highest No Observed Effect Concentration
IR	Ingestion Rate
ke	Rate of selenium loss
ku	Rate of selenium uptake
LOEC	Lowest Observed Effect Concentration
M	Muscle
MATC	Maximum Acceptable Toxicant Concentration (expressed mathematically as the
geometric mean of the NOEC and LOEC)
MDR	Minimum Data Recommendations or Requirements
NPDES	National Pollutant Discharge Elimination System
NOEC	No Observed Effect Concentration
SMCV	Species Mean Chronic Value
SSD	Species Sensitivity Distribution
TMDL	Total Maximum Daily Load
TRAP	EPA's Statistical Program: Toxicity Relationship Analysis Program
TTF	Trophic Transfer Factor
WB	Whole body
WQBELS	Water Quality-based Effluent Limitations
WQC	Water Quality Criteria
WQS	Water Quality Standards
ww	Wet Weight
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Executive Summary
This document sets forth the basis for and derivation of the Clean Water Act, Section
304(a) water quality criterion for protecting freshwater aquatic life from harmful effects of
selenium, a naturally occurring chemical element that is nutritionally essential in small amounts,
but toxic at higher concentrations. This assessment provides a critical review of all data
identified in EPA's literature search quantifying the toxicity of selenium to freshwater aquatic
organisms, and provides a basis for a criterion that will assure protection of populations of fish,
amphibians, aquatic invertebrates, and plants, based on available data.
Although selenium may cause acute toxicity at high concentrations, the most deleterious
effect on aquatic organisms is due to its bioaccumulative properties; these chronic effects are
found at lower concentrations than acute effects. Organisms in aquatic environments exposed to
selenium accumulate it primarily through their diets, and not directly through water (Chapman et
al. 2010). The best science also indicates that selenium toxicity occurs primarily through transfer
to the eggs and subsequent reproductive effects. Consequently, in harmony with the
recommendations of expert panels (U.S. EPA 1998; Chapman et al. 2010) and with peer review
and public comments on previous U.S. EPA (2004, 2014, 2015) drafts, the Agency developed a
chronic criterion reflective of the reproductive effects of selenium concentrations on fish species.
The 2016 "Aquatic Life Ambient Water Quality Criterion for Selenium - Freshwater,
2016," is a chronic criterion that is composed of four elements. All elements are protective
against chronic selenium effects. Two elements are based on the concentration of selenium in
fish tissue and two elements are based on the concentration of selenium in the water column. The
recommended elements are: (1) a fish egg-ovary element; (2) a fish whole-body and/or muscle
element; (3) a water column element (one value for lentic and one value for lotic aquatic
systems); and (4) a water column intermittent element to account for potential chronic effects
from short-term exposures (one value for lentic and one value for lotic aquatic systems). The
assessment of the available data for fish, invertebrates, and amphibians indicates that a criterion
value derived from fish will protect the aquatic community. All four criterion elements applied
together should protect aquatic life from the chronic effects of exposure to total selenium in
waters inhabited by fish, as well as "fishless waters."
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Because the factors that determine selenium bioaccumulation vary among aquatic
systems, site-specific water column criterion element values may be necessary at aquatic sites
with high selenium bioaccumulation to ensure adequate protection of aquatic life (Appendix K).
Finally, this freshwater chronic selenium criterion applies only to aquatic life, and is not intended
to address selenium toxicity to aquatic-dependent wildlife such as aquatic-dependent birds.
The toxicity studies relevant to the derivation of the fish tissue selenium criterion
elements involve (a) extended duration dietary exposure, and (b) measurement of total selenium
in the tissue of the target organism. Selenium either in fish whole-body or in muscle is usually
measured in non-reproductive studies, and selenium in eggs or ovaries is typically measured in
reproductive studies. Selenium accumulation in the eggs of the exposed adult female prior to
spawning has been shown to yield the most robust relationship (statistically significant) with
occurrence of deformities and reduced survival of the offspring.
The outcome of assessing both reproductive and non-reproductive studies under
laboratory and field conditions led EPA to the conclusion, consistent with expert consensus
(Chapman et al. 2009, 2010), that reproductive effects, linked to egg-ovary selenium
concentrations, provide the most sound basis for the criterion compared to non-reproductive
(e.g., survivorship, growth) endpoints. Reproductive effects have been linked to observed
reductions in the populations of sensitive fish species in waterbodies having elevated
concentrations of selenium (Young et al. 2010). EPA applied the sensitivity distribution concepts
from the U.S. EPA Guidelines for Deriving Numerical National Water Quality Criteria for the
Protection of Aquatic Organisms and their Uses (Stephan et al. 1985) to derive the national
selenium criterion. Based on the available data, expressed as ECio values, the egg-ovary criterion
element concentration is 15.1 milligrams selenium per kilogram dry weight (mg Se/kg dw),
based primarily on 17 reproductive studies representing 10 fish genera.
EPA recommends states and tribes adopt all four elements of the criterion into their water
quality standards. Two elements are based on the concentration of selenium in fish tissue (eggs
or ovaries, and whole-body or muscle) and two elements are based on the concentration of
selenium in the water column (a 30-day chronic element and an intermittent exposure element).
Both water column elements are further refined into values for lentic waters (e.g., lakes and
impoundments) and lotic waters (e.g., rivers and streams). The difference between lentic and
lotic water column values reflect the observed difference in selenium bioaccumulation in these

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two categories of aquatic systems (ATSDR 2003; Luoma and Rainbow 2005; Orr et al. 2006;
Simmons and Wallschlagel 2005). EPA derived the intermittent exposure element based on the
chronic 30-day water column element and the fraction of any 30-day period during which
elevated selenium concentrations occur. EPA recommends the intermittent element to address
short-term exposures that contribute to chronic effects through selenium bioaccumulation (e.g.,
releases from storage ponds or other intermittent releases). EPA derived the values for the water-
column criterion elements from the egg-ovary element by assessing food-chain bioaccumulation
based on available data collected at lentic and lotic systems in the continental United States.
Thus, all four criterion elements are based on reproductive effects in freshwater fish.
EPA primarily used field studies in freshwater systems to provide quantitative estimates
of selenium bioaccumulation in particulate material (algae, detritus, and sediment) from water,
and used field observations and laboratory data to quantify and model the trophic transfer of
selenium from particulate material into invertebrates, and from invertebrates into fish. EPA
additionally used field and laboratory observations to assess species-specific selenium
partitioning between different tissues within a fish (whole-body, eggs and/or ovaries, and
muscle). EPA developed food web models of fish in aquatic systems with a range of
bioaccumulation potentials and used the food web models with the species-specific estimates of
trophic transfer (or the most proximate taxonomic surrogate when species-specific data was not
available) to develop water column criterion elements from the egg-ovary criterion element for
lotic and lentic aquatic systems. EPA validated this approach using selenium measurements from
aquatic systems with a range of bioaccumulation potentials. Similar approaches could be used in
the derivation of selenium criteria in saltwater or estuarine systems with selenium data and food
webs relevant to those systems.
While more than half the available chronic studies were fish studies, available field data
and laboratory toxicity studies suggest that a criterion based on fish will protect amphibians,
aquatic invertebrates, and plants since these taxa appear to be less sensitive to selenium than fish
(see Sections 3.1.4 and 6.1.4),
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Table 1. Summary of the Recommended Freshwater Selenium Ambient Chronic Water
Quality Criterion for Protection of Aquatic Life.	
Media
Type
I' isli Tissue1
\Yater Column4
Criterion
Klcmcnt
Kgg/Ovary 2
I'isli W hole
liody or
Muscle'
Monthly
Average
Kxposurc
Intermittent KxposuiV
Magnitude
1 5 I mg/kg dw
8.5 mg/kg dw
whole body
or
11.3 mg/kg
dw muscle
(skinless,
boneless filet)
1.5 |ig/L in
lentic aquatic
systems
3.1 |ig/L in lotic
aquatic systems
WQClnt =
WQCso-day ~ Cbkgrnd(.l ~ f int)
f int
Duration
Instantaneous
measurement6
Instantaneous
measurement6
30 days
Number of days/month with an
elevated concentration
frequency
Not to be
exceeded
Not to be
exceeded
Not more than
once in three
years on
average
Not more than once in three years on
average
1.	Fish tissue elements are expressed as steady-state.
2.	Egg/Ovary supersedes any whole-body, muscle, or water column element when fish egg/ovary concentrations are
measured, except as noted in footnote 4 below.
3.	Fish whole-body or muscle tissue supersedes water column element when both fish tissue and water concentrations are
measured, except as noted in footnote 4 below.
4.	Water column values are based on dissolved total selenium in water and are derived from fish tissue values via
bioaccumulation modeling. When selenium inputs are increasing, water column values are the applicable criterion element
in the absence of steady-state condition fish tissue data.
5.	Where WQC30-day is the water column monthly element, for either a lentic or lotic waters; Cbkgrnd is the average
background selenium concentration, and fint is the fraction of any 30-day period during which elevated selenium
concentrations occur, with fmt assigned a value >0.033 (corresponding to 1 day).
6.	Fish tissue data provide instantaneous point measurements that reflect integrative accumulation of selenium over time and
space in fish population(s) at a given site.
The recommended chronic selenium criterion is expected to protect the entire aquatic
community, including fish, amphibians, and invertebrates, based on available data. Because fish
are the most sensitive to selenium effects, EPA recommends that selenium water quality criterion
elements based on fish tissue (egg-ovary, whole body, and/or muscle) data take precedence over
the criterion elements based on water column selenium data due to the fact, noted above, that fish
tissue concentrations provide a more robust and direct indication of potential selenium effects in
fish. However, because selenium concentrations in fish tissue are a result of selenium
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bioaccumulation via dietary exposure, there are two specific circumstances where the fish tissue
concentrations do not fully represent potential effects on fish and the aquatic ecosystem: 1)
"Ashless" waters, and 2) areas with new selenium inputs.
For purposes of this document, EPA defines "Ashless waters" as waters with insufficient
instream habitat and/or flow to support a population of any fish species on a continuing basis, or
waters that once supported populations of one or more fish species but no longer support fish
(e.g., extirpation) due to temporary or permanent changes in water quality (e.g., selenium
pollution), flow or instream habitat. Because of the inability to collect sufficient fish tissue to
measure selenium concentrations in fish in such waters, water column concentrations will best
represent selenium levels required to protect aquatic communities and downstream waters in
such areas. Appendix K of this criterion document discusses approaches to develop a site-
specific water column criterion element in such situations.
For purposes of this document EPA defines "new inputs" as new activities resulting in
the release of selenium into a lentic or lotic aquatic system. New inputs will likely result in a
greater concentrations of selenium in the food web and a relatively slow increase in the selenium
concentration in fish until the new selenium release achieves a quasi-"steady-state" balance in
the aquatic system. EPA estimates that the concentration of selenium in fish tissue will not reach
steady state for several months in lotic systems and longer time periods (e.g., 2 to 3 years) in
lentic systems. Achievement of steady state in an aquatic system also depends on the
hydrodynamics of the aquatic system, (particularly reservoirs with multiple riverine inputs), the
location of the selenium input and the particular food web. EPA expects the time needed to
achieve steady state with new or increased selenium inputs to be site specific. Thus, EPA
recommends that fish tissue criterion elements not take precedence over the water column
criterion elements until the aquatic system achieves steady state. In the interim, EPA
recommends sampling and using site-specific data to determine steady state in the receiving
water to gain a better understanding of the selenium bioaccumulation dynamics in a given
system.
EPA recommends states and tribes adopt into their water quality standards a selenium
criterion that expresses the four elements as a single criterion composed of multiple parts in a
manner that explicitly affirms the primacy of the whole-body or muscle element over the water
column elements, and the egg-ovary element over any other element. Adopting the fish whole-
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body or muscle tissue element into water quality standards ensures the protection of aquatic life
when measurements from fish eggs or ovary are not available, and adopting the water column
element ensures protection when fish tissue measurements are not available.
EPA recommends that when states adopt a four-part criterion for selenium reflecting
EPA's recommended criterion, states use the default monthly average exposure water column
elements of the criterion, adopted as part of the state's water quality criterion when implementing
the criterion under the National Pollutant Discharge Elimination System (NPDES) permits
program and to assist with implementation of other Clean Water Act programs. Alternatively,
states may want to develop adopt, and submit for EPA approval, either a site-specific water
column criterion element (or set of lentic/lotic criterion element values), or a set of procedures to
facilitate the translation of the fish tissue criterion concentration elements into site-specific water
concentration values. A site-specific water column criterion element or set of lentic/lotic
criterion element values can be developed using a mechanistic modeling approach (Presser and
Luoma 2010) or using the empirical bioaccumulation factor approach, both described in
Appendix K, for the specific waterbody or waterbodies. Any translation procedure must be
scientifically defensible, produce repeatable, predictable outcomes, and result in criterion
element values that protect the applicable designated use. Examples of such procedures include
the mechanistic modeling approach and the empirical BAF approach described in Appendix K.
This recommended selenium criterion applies to freshwater lentic and lotic systems, as it
is based on the toxicity of selenium to freshwater organisms. A similar approach may be
appropriate for deriving criteria for selenium in estuarine and marine waters if appropriate data
are available.
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1 Introduction and Background
The objective of the Clean Water Act (CWA) is to "restore and maintain the chemical,
biological and physical integrity of the Nation's waters." One of the tools that EPA uses to meet
this objective is the development of recommended ambient water quality criteria (AWQC) under
section 304(a)(1) of the Act. As provided for by the Clean Water Act, EPA reviews and from
time to time revises 304(a)(1) AWQC to ensure the criteria are consistent with the latest
scientific knowledge. Section 304(a) aquatic life criteria serve as recommendations to states and
authorized tribes for defining ambient water concentrations that will protect against adverse
ecological effects to aquatic life resulting from exposure to a pollutant found in water from direct
contact or, ingestion of contaminated water and/or food. Aquatic life criteria address the Clean
Water Act goals of providing for the protection and propagation of fish and shellfish. When
adopted into state or tribal water quality standards (WQS), these criteria can become a basis for
establishing National Pollutant Discharge Elimination System (NPDES) program permit limits
and, the basis for listing impaired waters under Section 303(d) and establishing Total Maximum
Daily Loads (TMDLs).
1.1 History of the EPA Recommended Selenium AWQC for Aquatic Life
In 1980 EPA first published numeric aquatic life criteria for selenium in freshwater.
These criteria were based on water-only exposure (no dietary exposure). In order to address the
lack of consideration of bioaccumulation in the 1980 selenium criteria, in 1987 EPA published
updated selenium criteria to address field-based toxicity observed in aquatic ecosystems at levels
below the existing criteria values. The 1987 criteria were field-based and accounted for both the
water column and dietary uptake pathways manifested at Belews Lake, North Carolina (USA), a
cooling water reservoir where water quality and fish communities had been affected by selenium
loads from a coal-fired power plant. At that time EPA also provided an acute criterion of 20 |ig/L
derived from a reverse application of an acute-chronic ratio obtained from conventional water-
only exposure toxicity tests applied to the 5 |ig/L chronic value based on dietary and water
column exposure in Belews Lake.
In 1998-1999 EPA published a revised acute criterion, a formula that recognized that the
two oxidation states, selenate and selenite, appeared to have substantially different acute
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toxicities. This acute criterion assumed toxicity was based on water-only exposure. Subsequent
research has demonstrated that sulfate levels influence selenate toxicity in water-only exposures.
In 1998 EPA held a peer consultation workshop (EPA-822-R-98-007) to evaluate new
science available for selenium relevant to the selenium aquatic life criterion. EPA concluded, and
the peer reviewers agreed, that fish-tissue values more directly represent chronic adverse effects
of selenium than the conventional water concentration approach used by EPA to protect aquatic
life, because chronic selenium toxicity is primarily based on the food-chain bioaccumulation
route, not on a water column route of exposure.
In 2004 EPA published a draft chronic whole-body fish-tissue criterion with a water-
based monitoring trigger in the summer and fall. The critical effect considered at that time was
the impact on survivorship based on overwintering stress to bluegill sunfish. An acute criterion
was estimated at that time that addressed concerns with the species of selenium present and
adjusted for sulfate levels; however, it did not address the dietary uptake pathway.
Further refinement of the fish tissue approach occurred in 2009 based on the findings of a
Pellston scientific workshop on the ecological risk assessment of selenium (Chapman et al. 2009,
2010). As presented by Chapman et al. (2009), some key findings resulting from that workshop
are:
	Diet is the primary pathway of selenium exposure for both invertebrates and vertebrates.
	Traditional methods for predicting toxicity on the basis of exposure to dissolved [water
column] concentrations do not work for selenium because the behavior and toxicity of
selenium in aquatic systems are highly dependent upon site-specific factors, including
food web structure and hydrology.
	Selenium toxicity is primarily manifested as reproductive impairment due to maternal
transfer, resulting in embryotoxicity and teratogenicity in egg-laying vertebrates.
In this 2016 final recommended freshwater chronic criterion for selenium, EPA includes
revisions based on the public and external expert peer reviews of the 2014 draft, public
comments on the 2015 draft, data and information from additional studies provided by the public
and peer reviewers, and additional scientific analyses. EPA also conducted a new literature
review and reanalyzed data considered in the 2004 and 2009 draft criteria documents. This final
criterion reflects the latest scientific consensus (e.g., Chapman et al. 2010) on the reproductive
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effects of selenium on aquatic life and their measurement in aquatic systems and supersedes all
previous national aquatic life water quality criteria for selenium.
EPA is recommending a national selenium criterion expressed as four elements. All
elements are protective against chronic selenium effects, and account for both short term and
longer term exposure to selenium. Two elements are based on the concentration of selenium in
fish tissue (eggs and ovaries, and whole-body or muscle) and two elements are based on the
concentration of selenium in the water-column (two 30-day chronic values and an intermittent
value). EPA derived the 30-day chronic water column element from the egg-ovary element by
modeling selenium bioaccumulation in food webs of lotic and lentic aquatic systems. EPA is
recommending the intermittent value to address short-term exposures that could contribute to
chronic effects through selenium bioaccumulation in either lotic or lentic systems. EPA derived
the intermittent element based on the chronic 30-day water column element and the fraction of
any 30-day period during which elevated selenium concentrations occur. These water column
criterion elements apply to the total of all oxidation states (selenite, selenate, organic selenium,
and any other forms) (See Appendix L for Analytical Methods for Measuring Selenium). Aquatic
communities are expected to be protected by this chronic criterion from any potential acute
effects of selenium.
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2 Problem Formulation
Problem formulation provides a strategic framework for water quality criteria
development by focusing the effects assessment on the most relevant chemical properties and
endpoints. The structure of this effects assessment is consistent with EPA's Guidelines for
Ecological Risk Assessment (U.S. EPA 1998).
This ecological effects assessment defines a scientifically-defensible water quality
criterion for selenium under section 304(a)(1) of the Clean Water Act. Clean Water Act Section
304(a)(1) requires EPA to develop criteria for water quality that accurately reflect the latest
scientific knowledge. These criteria are based solely on high quality data and best professional
scientific judgments on toxicological effects. Criteria are developed following overarching
guidance outlined in the Agency's Guidelines for Deriving Numerical National Water Quality
Criteria for the Protection of Aquatic Organisms and Their Uses (Stephan et al. 1985), hereafter
referred to as "EPA Ambient Water Quality Criteria Guidelines." States and authorized tribes
may adopt EPA's recommended criteria into their water quality standards to protect designated
uses of water bodies, they may modify EPA's criteria to reflect site-specific conditions, or they
may derive criteria using other scientifically-defensible methods, all subject to EPA review and
approval.
2.1 Overview of Selenium Sources and Occurrence
Selenium is a naturally occurring element present in sedimentary rocks and soils. It is
also present in the aquatic environment as methyl derivatives of selenium, naturally occurring in
freshwaters through methylation by bacteria (Ranjard et.al. 2003). Selenium's occurrence in
surficial soils and aquatic sediments in the United States is illustrated in Figure 2.1. There are
around 40 known selenium-containing minerals, some of which can have as much as 30%
selenium, but all are rare and generally occur together with sulfides of metals such as copper,
zinc and lead (Emsley 2011). Sedimentary rocks, particularly shales, have the highest naturally
occurring selenium content (Burau 1985). The distribution of organic-enriched, sedimentary
shales, petroleum source rocks, ore deposits, phosphorites, and coals, in which selenium
typically co-occurs, is well characterized in the United States (Presser et al. 2004). Natural
weathering of selenium-bearing geologic strata containing selenium can lead to selenium
leaching into groundwater and surface water. Two major anthropogenic activities are known to
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cause increased selenium mobilization and introduction into aquatic systems. The first is the
mining of metals, minerals and refinement and use of fossil fuels; the second is irrigation of
selenium-rich soils.
Mining activities bring selenium-enriched deposits to the surface, where they are exposed
to physical weathering processes. The release of selenium related to resource extraction activities
is most common in the phosphate deposits of southeast Idaho and adjacent areas of Wyoming,
Montana, and Utah, and in coal mining areas in portions of West Virginia, Kentucky, Virginia,
and Tennessee (Presser et al. 2004). Where selenium-containing minerals, rocks, and coal are
mined, selenium can be mobilized when rock overburden and waste materials are crushed,
increasing the surface area and exposure of material to weathering processes. Selenium
contamination of surface waters can also occur when sulfide deposits of iron, uranium, copper,
lead, mercury, silver, and zinc are released during the mining and smelting of these metal ores.
Where coal is burned for power production, selenium can enter surface waters as drainage from
fly-ash ponds and fly-ash deposits on land (Gillespie and Baumann 1986). Fly ash deposits have
a high surface area to volume ratio, resulting in rates of selenium in leachate several times higher
than from the parent feed coal (Fernandez-Turiel et al. 1994). The refining of crude oil
containing high levels of selenium can also be a major source of loading in certain water bodies
(Maher et al. 2010).
Irrigation of selenium-rich soils for crop production in arid and semi-arid regions of the
country can mobilize selenium and move it off-site in surface water runoff or via leaching into
ground water. Where deposits of Cretaceous marine shales occur, they can weather to produce
high selenium soils; such soils are present in many areas of the western U.S. (Lemly 1993c).
Selenium is abundant in the alkaline soils of the Great Plains, and some ground waters in
California, Colorado, Kansas, Oklahoma, South Dakota and Wyoming contain elevated
concentrations of selenium due to weathering of and leaching from rocks and soils. In semi-arid
areas of the West, irrigation water applied to soils containing soluble selenium can leach
selenium. The excess water (in tile drains or irrigation return flow) containing selenium can be
discharged into basins, ponds, or streams. For example, elevated selenium levels at the Kesterson
Reservoir in California originated from agricultural irrigation return flow collected in tile drains
that discharged into the reservoir (Ohlendorf et al. 1986).
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Se (ppm)
No data
0.01-0.12
0.13-0.14
0.14- 0.15
0.15-0.17
0.17-0.19
0.19-0.20
0.20-0.22
0.22-0.24
0.24-0.25
0.25- 0.27
0.27-0.29
0.29-0.32
0.32-0.35
0.35-0.38
0.38-0.41
0.41 - 0.45
0.45-0.51
0.51 - 0.59
0.59-0.75
0.75-5.32
Figure 2.1. Selenium in Surficial Soils and Aquatic Sediments in counties of the
Conterminous United States.
U.S. Geological Survey Open-File Report 2004-1001. URL:
http://mrdata.usgs.gov/geochem/doc/averages/countvdata.htm. Data are available from:
http://mrdata.usgs.gov/geochem/doc/groups-cats.htm.
Atmospheric emissions of selenium can originate from several sources including power
plants and other facilities that burn coal or oil, selenium refineries that provide selenium to
industrial users, base metal smelters and refineries, resource extraction industries, milling
operations, and end-product manufacturers (e.g., semiconductor manufacturers) (ATSDR 2003).
Airborne selenium particles can settle either on surface waters or on soils from which selenium
can be further transported and deposited into water bodies through ground or surface water
conveyances or runoff.
The chemical form of selenium that dominates a location is usually dependent on its
sources, effluent treatments, and biogeochemical processes in the receiving waters. Irrigation
2_
activities in areas with seleniferous soils typically mobilize selenate (SeC>4 , or Se[VI]) (Seiler et
2	
al. 2003). Combustion of coal for power generation creates predominantly selenite (SeCb , or
Se[IV|) in the fly ash waste due to the temperatures, pFI, and redox conditions involved with the
6

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process (Huggins et al. 2007). Similar conditions during refinement of crude oil can also result in
high concentrations of selenite relative to selenate, as was observed in the San Francisco Bay
estuary in the 1980s (Cutter 1989). Although selenite is the dominant species in the discharges
resulting from crude oil refining and coal burning using conventional technologies, the
implementation of alternative treatment technologies can alter the relative concentrations of
selenate and selenite. For example, in scrubbers with forced oxidation systems that produce
strong oxidizing conditions and high temperatures, the majority of discharged selenium is in the
form of selenate (Maher et al. 2010). However for flue gas desulfurization systems that are the
inhibited oxidation type, the selenium chemistry is more complex, and selenite may not be the
primary form emitted (Petrov et al. 2012). Table 2.1 shows the predominant form of selenium
that is associated with different activities and industries.
EPA's Office of Water and Office of Research and Development conducted the first
statistically based survey of contaminants in fish fillets from U.S. rivers from 2008 through
2009. This national fish survey was conducted under the framework of EPA's National Rivers
and Streams Assessment (NRSA), a probability-based survey designed to assess the condition of
the Nation's streams and rivers (Lazorchak et al. 2014). During June through October of 2008
and 2009, field teams applied consistent methods nationwide to collect samples of fish species
commonly consumed by humans at 541 randomly selected river locations (> 5th order based on
l:100,000-scale Strahler order) in the lower 48 states. They collected one composite fish sample
at every sampling location, with each composite consisting of five similarly sized adult fish of
the same species from a list of target species. Largemouth and smallmouth bass were the primary
species collected for the study, accounting for 34% and 24% of all fish composites, respectively.
Samples were collected from both non-urban (379 sites) and urban locations (162 sites). Each
fillet composite sample was homogenized and analyzed using an ICP-MS (Inductively Coupled
Plasma- Mass Spectrometry) method for total selenium, and results were reported as wet weight.
Three of the 541 samples (approximately 0.6%) exceeded the 2016 criterion for muscle tissue,
11.3 mg/kg dw. The maximum value detected was 17.75 mg Se/kg dw muscle, the median was
1.90 mg Se/kg dw, and the minimum 0.41 mg Se/kg dw.
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Table 2.1. Predominant Chemical Forms of Selenium in Discharges Associated with
Different Activities and Industries.
Selenium l-orm
Sources
Selenate
Agricultural irrigation drainage
Treated oil refinery effluent
Mountaintop coal mining/ valley fill leachate
Copper mining discharge
Selenite
Oil refinery effluent
Fly ash disposal effluent
Phosphate mining overburden leachate
Organoselenium
Treated agricultural drainage (in ponds or lagoons)
Source: Presser and Ohlendorf 1987; Zhang and Moore 1996; Cutter and Diego-McGlone 1990.
2.2 Environmental Fate and Transport of Selenium in the Aquatic
Environment
The fate and transport of selenium in aquatic systems is affected by the distribution of
selenium species and their transformations in water, sediment, and biota. These transformations
include the assimilation and conversion of inorganic selenium to organic selenium species in
plants and microbes that are transferred to higher trophic level consumer species throughout the
aquatic food web.
2.2.1 Selenium Species in Aquatic Systems
Aquatic organisms are exposed to a combination of predominantly organic selenium
species present in the food web throughout their life history; reproductive effects integrate these
exposures to transformed inorganic and organic species of selenium. The bioavailability and
toxicity of selenium depend on both its concentration and speciation (Cutter and Cutter 2004;
Meseck and Cutter 2006; Reidel et al. 1996). Selenium exists in four oxidation states (VI, IV, 0, -
II) and in a wide range of chemical and physical species across these oxidation states (Doblin et
al. 2006; Maher et al. 2010; Meseck and Cutter 2006). Therefore, in the effects assessment that
follows, we have correlated the adverse effects on aquatic life with total dissolved selenium.
In oxygenated surface waters, the primary dissolved selenium species are selenate
9_	9_
(SeC>4 , or Se[VI]) and selenite (SeC>3 , or Se[IV]), as well as dissolved organic selenides (-II)
formed from fine particulate organic matter (e.g., Doblin et al. 2006; Meseck and Cutter 2006).
The relative abundance of selenate and selenite depends on relative contributions from the
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geologic and anthropogenic sources of selenium to the receiving waters, as there is negligible
inter-conversion between the two species (e.g., Maher et al. 2010). Aqueous selenite is more
abundant than selenate when the majority of selenium originates from discharges from coal fly
ash tailings or oil refineries (e.g., Cutter 1989; Huggins et al. 2007). Particulate species in the
water column include selenate, selenite, and elemental selenium (Se(0)) bound to resuspended
sediments and organic particles, as well as particulate organic selenium species incorporated into
suspended detritus (e.g., Cutter and Bruland 1984; Meseck and Cutter 2006).
In sediments, selenate and selenite can be reduced to iron selenides or elemental selenium
under abiotic or biotic processes; elemental selenium and selenides can be converted to selenate
under oxidizing conditions (Maher et al. 2010). For example, selenate can be reduced to
elemental selenium in sediments (e.g., Oremland 1990) in the presence of iron oxides (Chen et
al. 2008) and iron sulfides (Breynaert et al. 2008). Elemental selenium and organic selenides are
produced by selenate-reducing microbes in sediments. Overall, the reduction of selenate and
particularly selenite in sediments increases with increasing sediment organic matter (Tokunaga et
al. 1997). Selenite in particular is readily bound to iron and manganese oxy-hydroxides (Maher
et al. 2010), and is readily adsorbed to inorganic and organic particles, particularly at a lower pH
range (e.g., McLean and Bledsoe 1992; Tokungawa et al. 1997). Microbial reduction of selenite
to organic forms (via methylation) increases the solubility and bioavailability of selenium
(Simmons and Wallschlagel 2005). Plants and algae produce volatile selenium species by
biomethylation of excess selenium, which upon reaching the sediment can be transformed to a
more bioavailable species, or deposited in the sediments and effectively removed from the
system (Diaz et al. 2009). Depending on environmental conditions, the reduction processes
described above are largely reversible, as elemental selenium and selenides in sediments can be
oxidized to selenate through microbial or abiotic transformations (e.g., Maher et al. 2010;
Tokunaga et al. 1997).
The most important transformation of selenium, with respect to its toxicity to aquatic
organisms, is in the uptake of dissolved inorganic selenium into the tissues of primary producers
at the base of the food web. The main route of entry of selenium into aquatic foodwebs is from
the consumption of particulate selenium of primary producers, and to a lesser degree, from the
consumption of sediments (Doblin et al. 2006; Luoma and Presser 2009). For algae, selenite and
organic selenides are similarly bioavailable, and both dissolved species are more bioavailable
9

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than selenate (e.g., Baines et al. 2001; Luoma et al. 1992). In vascular plants, selenate uptake is
greater than for the other dissolved species, as the majority of selenium uptake occurs in the
roots, and selenate is more easily transported to the shoots and leaves than selenite or organic
selenides (Dumont 2006). Following uptake, selenium is metabolized into a variety of organic
species that are assimilated into plant tissues. Selenium metabolism in plants is analogous to
sulfur metabolism (e.g., Dumont et al. 2006; Ouerdane et al. 2013). Selenate is reduced to
selenite, which is then reduced to selenide in a process involving reduced glutathione (Dumont et
al. 2006). Selenide is converted to selenocysteine (SeCys), which is then converted to
selenomethionine (SeMet) (Dumont et al. 2006). In addition to SeCys and SeMet, a variety of
other organic selenium species can be formed; however, SeCys, and particularly SeMet are
toxicologically important because these amino acids nonspecifically replace cysteine and
methionine in proteins, and are more bioavailable to higher trophic level consumers (Fan et al.
2002; Freeman et al. 2006).
2.2.2 Bioaccumulation of Selenium in Aquatic Systems
Dissolved selenium uptake by animals is slow, whatever the form, such that under
environmentally relevant conditions, dissolved selenium in the water column makes little or no
direct contribution to bioaccumulation in animals (Lemly 1985a; Ogle and Knight 1996), but
does influence the concentration of selenium in particulate matter. Selenium bioaccumulation in
aquatic organisms occurs primarily through the ingestion of food (Fan et al. 2002; Luoma et al.
1992; Maher et al. 2010; Ohlendorf et al. 1986; Presser and Ohlendorf 1987; Presser et al. 1994;
Saiki and Lowe 1987). However, unlike other bioaccumulative contaminants such as mercury,
the single largest step in selenium accumulation in aquatic environments occurs at the base of the
food web where algae and other microorganisms accumulate selenium from water by factors
ranging from several hundred to tens of thousands (Luoma and Presser 2009; Orr et al. 2012;
Stewart et al. 2010). Bioaccumulation and transfer through aquatic food webs are the major
biogeochemical pathways of selenium in aquatic ecosystems. Dissolved selenium oxyanions
(selenate, selenite) and organic selenides are assimilated into the tissues of aquatic primary
producers (trophic level 1 organisms), such as periphyton, phytoplankton, and vascular
macrophytes; and subsequently biotransformed into organoselenium. These organisms, together
with other particle-bound selenium sources, constitute the particulate selenium fraction in the
water column. Selenium from this particulate fraction is then transferred to aquatic primary
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consumers such as zooplankton, insect larvae, larval fish, and bivalves (trophic level 2), and then
to predators such as fish and birds (trophic level 3 and above). In addition to the water
concentration of selenium, the process of selenium bioaccumulation in aquatic life residing in
freshwater systems depends on several factors specific to each aquatic system. These factors
include:
Water residence time. Residence time is a measure of the average time a water molecule
will spend in a specified region of space. Residence time influences both the proportion of
selenium found in particulate and dissolved forms and the predominant form of selenium.
Organisms in waters with long residence times such as lakes, ponds, reservoirs, wetlands or
estuaries will tend to bioaccumulate more selenium than those living in waters with shorter
residence times such as rivers and streams (ATSDR 2003; EPRI2006; Luoma and Rainbow
2005; Orr et al. 2006; Simmons and Wallschlagel 2005). Several interrelated factors underlie
selenium's greater bioaccumulation potential in slow moving systems, such as food web
complexity and the organic content and reduction/oxidation potential of sediments. Finally,
selenium toxicity in flowing waters with shorter residence times may only be apparent
downstream of their selenium sources, whereas waters with longer residence times are more
likely to exhibit selenium toxicity near their sources (Presser and Luoma 2006).
Distribution of selenium between particulate and dissolvedforms. Selenium is found in
both particulate and dissolved forms in water, but direct transfer of selenium from water to
animals is only a small proportion of the total exposure. The proportion of selenium found in
particulate matter (algae, detritus, and sediment) is important because it is the primary avenue for
selenium entering into the aquatic food web (Luoma et al. 1992; Luoma and Rainbow 2005;
Ohlendorf et al. 1986; Presser and Ohlendorf 1987; Presser et al. 1994; Presser and Luoma 2006;
Saiki and Lowe 1987).
Bioaccumulation in prey. Trophic level 1 organisms such as periphyton and
phytoplankton, as well as other forms of particulate material containing selenium, such as
detritus and sediment, are ingested by trophic level 2 organisms such as mollusks, planktonic
crustaceans, and many insects, increasing the concentration of selenium in the tissues of these
higher-level organisms. Differences in the physiological characteristics of these organisms result
in different levels of bioaccumulation. Also, selenium effects on invertebrates typically occur at
concentrations higher than those that elicit effects on the vertebrates (e.g., fish and birds) that
11

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prey upon them. Additionally, certain molluscan taxa such as mussels and clams can accumulate
selenium to a much greater extent than planktonic crustaceans and insects (although the levels do
not seem to be toxic to the mussels) due to higher ingestion rates of both particulate-bound
(algae) and dissolved selenium from the water column through filter feeding, as well as the lower
rate at which they eliminate selenium (Luoma and Rainbow 2005; Stewart et al. 2013). Because
egg-laying (oviparous) vertebrates such as fish and birds are most sensitive to selenium effects,
(Janz et al. 2010), these vertebrate consumers are also the most vulnerable groups to selenium
poisoning and the focal point of most selenium environmental assessments (Ogle and Knight
1996; Stewart et al. 2010).
Trophic transfer to predators. Bioaccumulation of selenium by higher trophic level
organisms, such as trophic level 3 and 4 fish, is highly influenced by the food web of the aquatic
environment. For example, fish that primarily consume freshwater mollusks (e.g., redear sunfish)
will exhibit greater selenium bioaccumulation than fish that consume primarily insects or
crustaceans from waters with the same concentration of dissolved selenium because mollusks
tend to accumulate selenium at higher concentrations than other trophic level 2 organisms, as
noted above (Luoma and Presser 2009; Stewart et al. 2004).
2.3 Mode of Action and Toxicity of Selenium
Selenium is a naturally occurring chemical element that is also an essential micronutrient.
Trace amounts of selenium are required for normal cellular function in almost all animals.
However, excessive amounts of selenium can also have toxic effects, with selenium being one of
the most toxic of the biologically essential elements (Chapman et al. 2010). Egg-laying
vertebrates have a lower tolerance than do mammals, and the transition from levels of selenium
that are biologically essential to those that are toxic occurs across a relatively narrow range of
exposure concentrations (Luckey and Venugopal 1977; U.S. EPA 1987, 1998; Haygarth 1994;
Chapman et al. 2009, 2010). Selenium consumed in the diet of adult female fish is deposited in
the eggs, when selenium replaces sulfur in vitellogenin, which is transported to the ovary and
incorporated into the developing ovarian follicle (Janz et al. 2010), the primary yolk precursor.
Selenium is a member of the sulfur group of nonmetallic elements, and consequently, the
two chemicals share similar characteristics. Selenium can replace sulfur in two amino acids, the
seleno-forms being selenomethionine and selenocysteine. It has been a long-standing hypothesis
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that the cause of malformations in egg-laying vertebrates is due to the substitution of selenium
for sulfur in these amino acids and their subsequent incorporation into proteins, which causes
disruption of the structure and function of the protein. When present in excessive amounts,
selenium is erroneously substituted for sulfur, resulting in the formation of a triselenium linkage
(Se-Se-Se) or a selenotrisulfide linkage (S-Se-S), either of which was thought to prevent the
formation of the normal disulfide chemical bonds (S-S). The end result was thought to be
distorted, dysfunctional enzymes and protein molecules that impaired normal cellular
biochemistry (Diplock and Hoekstra 1976; Reddy and Massaro 1983; Sunde 1984).
Recent research, however, suggests that selenium's role in oxidative stress plays a role in
embryo toxicity, whereas selenium substitution for sulfur does not. The substitution of
selenomethionine for methionine does not appear to affect either the structure or function of
proteins (Yuan et al. 1998; Mechaly et al. 2000; Egerer-Sieber et al. 2006). The reason is
apparently due to selenium not being distally located in selenomethionine, which insulates the
protein from an effect on its tertiary structure. Although the incorporation of selenomethionine
into proteins is concentration-dependent (Schrauzer 2000), selenocysteine's incorporation into
proteins is not (Stadtman 1996). This suggests that neither selenomethionine nor selenocysteine
affect protein structure or function. In fact, Se as an essential micronutrient is incorporated into
functional and structural proteins as selenocysteine.
The role of selenium-induced oxidative stress in embryo toxicity and teratogenesis
appears to be related to glutathione homeostasis. A review of bird studies by Hoffman (2002)
showed exposure to selenium altered concentrations and ratios of reduced to oxidized glutathione
thereby increasing measurements of oxidative cell damage. Palace et al. (2004) suggested
oxidative stress due to elevated selenium levels results in pericardial and yolk sac edema in
rainbow trout embryos. Evidence for the role of oxidative stress in selenium toxicity is growing,
but mechanistic studies are needed to better understand its effects on egg-laying vertebrates. For
a more in depth discussion on the mechanism of toxicity at the cellular level including the
evidence against sulfur substitution as a cause and the role of oxidative stress, see Janz et al.
(2010).
The most well-documented, overt and severe toxic symptoms in fish are reproductive
teratogenesis and larval mortality. Egg-laying vertebrates appear to be the most sensitive taxa,
with toxicity resulting from maternal transfer to eggs. In studies involving young organisms
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exposed through transfer of selenium from adult female fish into their eggs, the most sensitive
diagnostic indicators of selenium toxicity in vertebrates occur when developing embryos
metabolize organic selenium that is present in egg albumen or yolk. It is then further metabolized
by larval fish after hatching.
A variety of lethal and sublethal deformities can occur in the developing fish exposed to
selenium, affecting both hard and soft tissues (Lemly 1993b). Developmental malformations are
among the most conspicuous and diagnostic symptoms of chronic selenium poisoning in fish.
Terata are permanent biomarkers of toxicity, and have been used to identify impacts of selenium
on fish populations (Maier and Knightl994; Lemly 1997b). Deformities in fish that affect
feeding or respiration can be lethal shortly after hatching. Terata that are not directly lethal, but
distort the spine and fins, can reduce swimming ability, and overall fitness. Because the rate of
survival of deformed young would be less than that for normal young, the percentage of
deformed adults observed during biosurveys will likely understate the underlying percentage of
deformed young, although quantitation of the difference is ordinarily not possible.
In summary, the most sensitive indicators of selenium toxicity in fish larvae are effects
modulated through the reproductive process and exhibited in fish larvae as teratogenic
deformities such as skeletal, craniofacial, and fin deformities, and various forms of edema that
result in mortality (Lemly 2002). The toxic effect generally evaluated is the reduction in the
number of normal healthy offspring compared to the starting number of eggs. In studies of young
organisms exposed to selenium solely through their own diet (rather than via maternal transfer),
reductions in survival and/or growth are the effects that are generally evaluated.
2.4 Narrow Margin between Sufficiency and Toxicity of Selenium
Selenium has a narrow range encompassing what is beneficial for biota and what is
detrimental. Selenium is an essential nutrient that is incorporated into functional and structural
proteins as selenocysteine and selenomethionine. Several of these proteins are enzymes that
provide cellular antioxidant protection. Selenomethionine is readily oxidized, and its antioxidant
activity arises from its ability to deplete reactive oxygen species. Selenomethionine is required as
a mineral cofactor in the biosynthesis of glutathione peroxidases. All of the classic glutathione
peroxidases contain selenium and are found to be involved in the catalytic reaction of these many
enzymes (Allan 1999). The major function of the glutathione peroxidases involves the reduction
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of hydrogen peroxide to water at the expense of the oxidation of glutathione, the enzyme's
cofactor, an important antioxidant process at normal dietary levels.
Aquatic and terrestrial organisms require low levels of selenium in their diet to sustain
metabolic processes, whereas excess concentrations of selenium that are only an order of
magnitude greater than the required level have been shown to be toxic to fish, apparently due to
generation of reactive oxidized species, resulting in oxidative stress (Palace et.al. 2004). Dietary
requirements in fish have been reported to range from 0.05 to 1.0 mg Se/kg dw (Watanabe et al.
1997). Selenium requirements for optimum growth and liver glutathione peroxidase activity in
channel catfish were reported as 0.25 mg Se/kg dw (Gatlin and Wilson 1984). Estimated
selenium dietary requirements in hybrids of striped bass, based on selenium retention, were
reported as 0.1 mg Se/kg dw (Jaramillo 2006). Selenium deficiency has been found to affect
humans (U.S. EPA 1987), sheep and cattle (U.S. EPA 1987), deer (Oliver et al. 1990), fish
(Thorarinsson et al. 1994; Wang and Lovell 1997; Wilson et al. 1997; U.S. EPA 1987), aquatic
invertebrates (Audas et al. 1995; Caffrey 1989; Cooney et al. 1992; Cowgill 1987; Cowgill and
Milazzo 1989; Elendt 1990; Elendt and Bais 1990; Harrison et al. 1988; Hyne et al. 1993;
Keating and Caffrey 1989; Larsen and Bjerregaard 1995; Lim and Akiyama 1995; Lindstrom
1991; U.S. EPA 1987; Winner 1989; Winner and Whitford 1987), and algae (Doucette et al.
1987; Keller et al. 1987; Price 1987; Price et al. 1987; Thompson and Hosja 1996; U.S. EPA
1987; Wehr and Brown 1985). The predominance of research on selenium deficiency in
invertebrates and algae is related to optimizing the health of test organisms cultured in the
laboratory. A summary of several studies that evaluated the deficiency and/or the sufficiency of
selenium in the diet of fish is provided in Appendix E.
2.5 Interactions with Mercury
The most well-known interactions with selenium occur with both inorganic and organic
mercury, and are generally antagonistic (Micallef and Tyler 1987; Cuvin and Furness 1988;
Paulsson and Lundbergh 1991; Siegel et al. 1991; Southworth et al. 1994; Ralston et al. 2006),
with the most likely mechanism being the formation of metabolically inert mercury selenides
(Ralston et al. 2006; Peterson et al. 2009). However, other studies have found interactions
between mercury and selenium to be additive (Heinz and Hoffman 1998) or synergistic
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(Huckabee and Griffith 1974; Birge et al. 1979). The underlying mechanism for these additive
and synergistic interactions between mercury and selenium are unknown.
2.6 Assessment Endpoints
Assessment endpoints are defined as "explicit expressions of the actual environmental
value that is to be protected" and are defined by an ecological entity (species, community, or
other entity) and its attribute or characteristics (U.S. EPA 1998). Assessment endpoints may be
identified at any level of organization (e.g., individual, population, community). In the context of
the Clean Water Act, aquatic life criteria for toxic pollutants are typically determined based on
the results of toxicity tests with aquatic organisms in which unacceptable effects on growth,
reproduction, or survival occurred. The goal of criteria is to protect the diversity, productivity,
and stability of aquatic communities. To achieve this goal, the endpoint of criteria assessment is
the survival, growth, and reproduction of a high percentage of species of a diverse assemblage of
freshwater aquatic animals (fish, amphibians, and invertebrates) and plants. Toxicity data are
aggregated into a sensitivity distribution that indicates the impact of the toxicant under study to a
variety of genera representing the broader aquatic community. Criteria are designed to be
protective of the vast majority of aquatic animal species in an aquatic community (i.e.,
approximately 95th percentile of tested aquatic animals representing the aquatic community). As
a result, health of the aquatic community may be considered as an assessment endpoint indicated
by survival, growth, and reproduction. Assessment endpoints are the ultimate focus in risk
characterization and link the measurement endpoints to the risk management process (e.g., policy
goals). When an assessment endpoint can be directly measured, the measurement and assessment
endpoints are the same. In most cases, however, the assessment endpoint cannot be directly
measured, so a measurement endpoint (or a suite of measurement endpoints) is selected that can
be related, either qualitatively or quantitatively, to the assessment endpoint. For example, a
decline in a sport fish population (the assessment endpoint) may be evaluated using laboratory
studies on the mortality of surrogate species, such as the fathead minnow (the measurement
endpoint) (EPA/630/R-92/001 February 1992). The assessment endpoint for selenium is the
protection of freshwater aquatic life; because we know that fish are the most sensitive aquatic
taxon to the toxicological effect of selenium, the criterion is expressed in terms of fish tissue
using eggs and ovarian tissue as the most representative element related to selenium toxicity.
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To assess potential effects on the aquatic ecosystem by a particular stressor, and develop
304(a) aquatic life criteria under the CWA, EPA typically requires the following, as outlined in
the EPA Ambient Water Quality Criteria Guidelines: acute toxicity test data (mortality,
immobility, loss of equilibrium) for aquatic animals from a minimum of eight diverse taxonomic
groups; as well as chronic toxicity data (e.g., survival, growth and reproduction) for aquatic
animals from 8 eight taxonomic groups (described in more detail below). The diversity of tested
species is intended to ensure protection of various components of an aquatic ecosystem. In the
case of bioaccumulative compounds like selenium, these acute toxicity studies do not address
risks that result from exposure to chemicals via the diet (through the food web). They also do not
account for the slow accumulation kinetics of many bioaccumulative compounds such as
selenium and may underestimate effects from long-term accumulation in different types of
aquatic systems (SAB 2005).
Because the most sensitive adverse effects of selenium are reproductive effects (larval
deformities and mortality) on the offspring of exposed fish, chronic effects from long term
exposure are the focus of this selenium assessment. In addition to continuous discharges, shorter-
term intermittent or pulsed exposures to elevated levels of selenium may also result in
bioaccumulation through the aquatic food web and may subsequently adversely affect fish
reproduction, and such measures of effect are therefore also estimated from chronic assessment
endpoints. Selenium toxicity in the water body could potentially threaten fecundity and
recruitment in fishes, resulting in extirpation of sensitive species in a waterbody, and potentially
shifting the trophic dynamics of the system. Therefore, the assessment endpoint for selenium is
the protection of fish populations. In some waters where ESA-listed fish species occur, a
protection goal oriented to protection of individuals may be more appropriate. This should be
reflected using site-specific data to derive an SSC for the site.
Chronic toxicity test data (longer-term survival, growth, or reproduction) for aquatic
animals are needed from a minimum of eight diverse taxonomic groups (or less generically,
[minimum of three taxa] if the derivation is based on an acute to chronic ratio). The diversity of
tested species is intended to ensure protection of various components of an aquatic ecosystem.
Specific minimum data recommendations or requirements (MDRs) identified for development of
criteria in the EPA Ambient Water Quality Criteria Guidelines require aquatic animal toxicity
data from:
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1.	the family Salmonidae in the class Osteichthyes,
2.	a second family in the class Osteichthyes, preferably a commercially or
recreationally important warmwater species (e.g., bluegill, channel catfish, etc.),
3.	a third family in the phylum Chordata (may be in the class Osteichthyes or may
be an amphibian, etc.),
4.	a planktonic crustacean (e.g., cladoceran, copepod, etc.),
5.	a benthic crustacean (e.g., ostracod, isopod, amphipod, crayfish, etc.),
6.	an insect (e.g., mayfly, dragonfly, damselfly, stonefly, caddisfly, mosquito,
midge, etc.),
7.	a family in a phylum other than Arthropoda or Chordata (e.g., Rotifera,
Annelida, Mollusca, etc.), and
8.	a family in any order of insect or any phylum not already represented.
Acceptable quantitative chronic values for selenium are available for six of the eight
MDRs (requirements 1, 2, 3, 6, 7, and 8). Acceptable chronic values for selenium are not
available for two of the MDRs (requirements 4 and 5: planktonic and benthic crustaceans,
respectively). Following the approach of U.S. EPA (2008b), which was reviewed by the Science
Advisory Board, if information is available to demonstrate that an MDR is not sensitive, then a
surrogate value can be used in place of actual toxicity data to represent the missing MDR. Based
on the data estimating the sensitivity of insects (Centroptilum triangulifer), rotifers (Brachionus
calyciflorus), and oligochaetes (Lumbriculus variegatus), EPA determined that invertebrates
(e.g., insects and crustaceans) are generally less sensitive to selenium than fish, based on the
characteristics of selenium toxicity to aquatic life. Therefore, the available fish data were used in
the genus-level sensitivity distribution to derive the chronic selenium criterion (Note:
invertebrate data were included in the sensitivity distribution for the whole body criterion
element to demonstrate that the derivation of the criterion element based on the fish egg-ovary to
whole body translated values protected invertebrates given the sensitivity range of the available
species).
The EPA Ambient Water Quality Criteria Guidelines also require at least one acceptable
test with a freshwater alga or vascular plant. If plants are among the aquatic organisms most
sensitive to the stressor, results of a plant in another phylum should also be available. A
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relatively large number of tests from acceptable studies of aquatic plants were available for
possible derivation of a Final Plant Value. However, the relative sensitivity of freshwater plants
to selenium (Appendix F) is less than fish, so plant criterion elements were not developed.
The available scientific evidence indicates that for selenium, critical assessment
endpoints for aquatic species are offspring mortality and severe development abnormalities that
affect the ability of fish to swim, feed and successfully avoid predation, resulting in impaired
recruitment of individuals into fish populations. Selenium enrichment of reservoir environments
(e.g., Belews Lake, NC (Lemly 1985), Hyco Reservoir (DeForest 1999), and Kesterson
Reservoir, CA (Ohlendorf 1986)) are well documented and demonstrate that adverse effects
resulted from bioaccumulative processes at different levels of biological organization, resulting
in population-level reductions of resident species.
2.7 Measures of Effect
Each assessment endpoint requires one or more "measures of ecological effect", which
are defined as changes in the attributes of an assessment endpoint itself or changes in a surrogate
entity or attribute in response to chemical exposure. Ecological effects data are used as measures
of direct and indirect effects to growth, reproduction, and survival of aquatic organisms.
The toxicity testing data available for any given pollutant vary significantly, depending
primarily on whether any major environmental issues are raised. An in-depth evaluation of
available data for selenium has been performed by EPA to determine data acceptability and
quality, based on criteria established in the EPA Ambient Water Quality Criteria Guidelines.
In traditional chronic tests used in many EPA aquatic life criteria documents, organisms
are exposed to contaminated water but fed a diet grown in uncontaminated media not spiked with
the toxicant prior to introduction into the exposure chambers. Such tests are not suitable for
deriving a criterion for a bioaccumulative pollutant unless (1) effects are linked to concentrations
measured in appropriate tissues, and (2) the route of exposure does not affect the potency of
residues in tissue. For selenium, the first condition might be met, but the second condition is not,
because the route of selenium exposure appears to influence the potency of a given tissue residue
(Cleveland et al. 1993; Gissel-Nielsen and Gissel-Nielsen 1978). Consequently, toxicity tests
with water-only exposures (and any tests not relying on dietary exposure) are not included in this
assessment.
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Selenium toxicity is primarily manifested as reproductive impairment due to maternal
transfer, resulting in embryo mortality and teratogenicity. Measurements of selenium in fish
tissue are most closely linked to the chronic adverse effects of selenium (Chapman et al. 2010),
since chronic selenium toxicity is based on the food-chain bioaccumulation route, not a direct
waterborne route. In this selenium criterion document, water column criterion element
concentrations for selenium were derived from fish tissue concentrations by modeling selenium
transfer through the food web. The next sections describe approaches used to establish selenium
effects concentrations in fish tissue and to relate the concentrations in fish tissue to
concentrations in water.
2.7.1 Fish Tissue
Chronic measures of effect concentrations are the ECio, EC20. No Observed Effect
Concentration (NOEC), Lowest Observed Effect Concentration (LOEC), and Maximum
Acceptable Toxicant Concentration (MATC). The EC10 is the concentration of a chemical that is
estimated to result in a 10 percent effect in a measured chronic endpoint (e.g., growth,
reproduction, or survival); the EC20 corresponds to 20 percent effect. The NOEC is the highest
chemical concentration at which none of the observed effects are statistically different from the
control, as determined by hypothesis testing. The LOEC is the lowest test concentration at which
observed effects are found to be statistically different from the control. For selenium, in all cases
the effect endpoint used in the estimation of chronic values (e.g., ECios) is an effect on offspring
(with exposure via maternal transfer) from parents exposed to selenium via diet.
Selenomethionine was used exclusively in dietary exposures in the lab, whereas field-exposed
females would be exposed to a combination of forms of selenium as a function of the selenium in
their prey items.
Whenever possible, estimates of selenium concentrations associated with a low level of
effect (i.e., EC10) were calculated for each study using the computer program TRAP (version
1.30a), Toxicity Relationship Analysis Program (U.S. EPA 2013). The program is based on a
regression approach that models the level of adverse effects as a function of increasing
concentrations of the toxic substance. With the fitted model it is possible to estimate the
contaminant concentration associated with a small effect. TRAP was used when there are
sufficient data for EC10 estimation. For studies with binary data, the analysis proceeded by
tolerance distribution analyses using the log-triangular distribution, unless there was substantial
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extrabinomial variability, in which case regression analysis was used. For regression analysis,
the threshold sigmoidal model was used, exposure variables were log-transformed, and effects
variables were weighted appropriately to address their relative uncertainties.
When there were insufficient data for TRAP to fit an effects/exposure curve (no
treatments with clear effects near the ECio and/or significant background variability), the ECio
was based on interpolation. To ensure that the interpolations were comparable to the TRAP
analyses, threshold sigmoidal equation was used. This equation is fitted to two points, and
constrained so that 3 equation parameters can be set. The first set-point was treated as the ECo
with a second associated set-point being the threshold for background effects values, based on
the highest NOEC (HNOEC) datum and other NOEC data. The final set-point was the LOEC. If
the LOEC is a partial effect, then this point was used to estimate the equation slope. If the LOEC
was a 100% effect, it was specified as the ECioo; with the ECo specified, then this relationship
dictated the equation slope. It should be noted that despite the superficial resemblance of these
analyses to TRAP they are also subject to the uncertainties associated with the interpolation
method.
It should be noted that TRAP involves the assumption that (a) there is a single underlying
relationship of the effects variable to the exposure variable which follows the specified equation
and (b) the exposure variable is known with negligible error, with uncertainty being
predominantly in the effects variable. Some of the reproductive data for selenium involved
multiple sources of variability that led to both multiple relationships across different cohorts of
offspring and uncertainty in the exposure variable, so that the resulting TRAP curves were more
approximate, and TRAP error estimates were generally not useful. These issues can also affect
the interpolation protocol. It should also be noted that estimating a concentration associated with
a low effects level, such as an ECio, is especially uncertain when treatments yielding partial
effects values are lacking in the concentration response data produced by a study. These two
issues prevented the use of TRAP in some datasets. When the data are insufficient to provide any
meaningful ECio by the first two approaches, the study should either not be used for criteria
development or a chronic value should be set by other means than an estimated ECio if possible.
Only studies with a reference site (field surveys) or control treatment(s) (experimental
studies) were included in the analysis, because response levels at these low selenium
concentrations were the most influential points for calculating the estimated response level at a
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selenium concentration of zero (yo). When considering the use of the ECio versus the EC20, an
EC10 was determined to be a more appropriate endpoint for tissue-based criteria given the nature
of exposure and effects for this bioaccumulative chemical. EC20s have historically been used in
the derivation of EPA criteria applicable to the water medium. While water concentrations may
vary rapidly over time, tissue concentrations of bioaccumulative chemicals are expected to vary
gradually over time. Thus, where concentrations of selenium in fish tissue are used as an effect
threshold, there is potential for sustained impacts on aquatic systems, relative to chemicals that
are not as bioaccumulative. Furthermore, it was found that the dose-response curves for selenium
across a broad range of fish genera are very steep, such that a small change in selenium tissue
concentration yielded a large increase in observed adverse effect. In many cases, the selenium
data indicated a change from control effect levels to effects in excess of 50% for larval mortality
or deformity over a few mg/kg dry weight increase in selenium detected in fish tissue. These
issues call for use of a lower level of effect to attain sufficient protection. The EC10 was also
preferred over the NOEC or LOEC as these measures of effect are influenced by study design,
specifically the concentrations tested, the number of concentrations tested, the number of
replicates for each concentration, and the number of organisms in each replicate. As noted by
Campbell (2011), EC10S and NOECs are generally of similar magnitude, but EC10S have the
advantage of being more reproducible than NOECs (Van der Hoeven et al. 1997; Warne and van
Dam 2008). NOECs and MATCs are generally presented if calculated by the original
investigators, but were not used where an EC10 could be calculated. The four lowest egg-ovary
Genus Mean Chronic Values (GMCVs), whose exact values influence the calculation of the egg-
ovary criterion element, are all based solely on EC10S. NOECs contribute to some of the GMCVs
for less sensitive species.
In this document, chronic values are presented as tissue concentrations of total selenium
in units of mg/kg dry weight (dw). Studies of chronic toxicity of selenium to aquatic organisms
measure concentrations in distinct tissues (e.g., whole body, ovaries, eggs, muscle, and liver) and
report these values as either wet weight (ww) or dw. Studies reporting tissue concentrations only
based on wet weight were converted to dry weight using tissue-specific and species-specific
conversion factors. When wet to dry weight conversion factors were not available for a given
species, conversion factors for a closely related taxon were used. In deriving the egg or ovary
tissue criterion element, chronic values are for those tissues directly measured in the study.
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Tissue-to-tissue conversions (e.g., to estimate concentrations in an unmeasured tissue from a
study's measured tissue) involve some uncertainty because of variability in tissue concentration
ratios (deBruyn et al. 2008; Osmudson et al. 2007). Tissue-to-tissue conversions were needed for
calculating the reproductive toxicity-based whole-body and muscle chronic criterion element and
water criterion concentration elements.
The overall assessment evaluates both reproductive and non-reproductive studies.
Selenium concentrations measured directly in eggs or ovaries from reproductive (maternal
transfer) studies are used to derive the egg/ovary criterion element, and corresponding selenium
concentrations in whole body or muscle tissue resulting in reproductive effects are estimated
using conversion factors. Direct measurements of selenium concentrations in whole-body or
muscle from non-reproductive studies are used to examine non-reproductive, chronic effects,
such as impairments to growth.
2.7.2 Water
While state monitoring programs may sample ambient waters for selenium, widespread
measurements of selenium in fish tissue are relatively rare. Therefore, EPA is providing
estimated chronic measures of effect for water column data. The chronic criterion element for the
water column is the 30-day average concentration that corresponds to the concentration of
selenium in fish tissue estimated to result in a 10 percent effect in fish for a specific water body
type (lotic or lentic water bodies as described below in Section 3.2.4), The chronic criterion
element for the water column is derived by modeling trophic transfer of selenium through the
food web resulting in the fish tissue concentration that yields the chronic reproductive effects of
concern.
EPA collaborated with the United States Geological Survey (USGS) to develop a model
(later published in Presser and Luoma 2010) that relates the concentration of selenium in fish
tissue to the water column. The approach is based on bioaccumulation and trophic transfer
through aquatic system food-webs. Model parameters are calculated using both field and
laboratory measurements of selenium in water, particulate material (algae, detritus and
sediment), invertebrates, fish whole-body, and fish egg-ovary. Although EPA and USGS use the
same model to relate the concentration of selenium in fish tissue to water, EPA starts with
selenium in the egg/ovary (reproductive criterion) whereas USGS starts with selenium in the
fish's whole body. The EPA approach therefore has the additional step of converting the
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concentration of selenium in the egg/ovary to whole body. This model (which is a set of
equations) is described in more detail in Section 3.2.1,
2.7.3 Summary of Assessment Endpoints and Measures of Effect
The typical assessment endpoints for aquatic life criteria are based on effects on growth,
reproduction, or survival of the assessed taxa. These measures of effect on toxicological
endpoints of consequence to populations are provided by results from toxicity tests with aquatic
plants and animals. The toxicity values (i.e., measures of effect expressed as genus means) are
used in the genus sensitivity distribution of the aquatic community to derive the aquatic life
criteria. Endpoints used in this assessment are listed in Table 2.2.
Table 2.2. Summary of Assessment Endpoints and Measures of Effect Used in Criterion
Derivation for Selenium.
Assessment Kndpoinls for (lie Aquatic
(ommunilY
Measures of KITecl
Survival, growth, and reproduction of
freshwater fish, other freshwater vertebrates,
and invertebrates
For effects from chronic exposure:
1.	ECio concentrations in egg and ovary, for
offspring mortality and deformity.
2.	Measured or estimated reproductive ECio
in whole body and muscle.
3.	Estimated concentrations (|ig/L) in water
linked to egg-ovary ECios by food web-
modeling.
4.	Intermittent water concentrations yielding
exposure equivalent to the above.
For acutely lethal effects:
Acute toxicity effects based on standard
water column-only toxicity testing are not
provided here for selenium, due to the
dominant significance of chronic effects.
Note: Chronic criterion is expected to be
protective of acute effects.
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2.7.4 Conceptual Model of Selenium Effects on Aquatic Life
Reproductive Impairment.
Larval skeletal deformities.
Larval mortality.
Tertiary Trophic Transfer from
lower trophic level fish
To higher trophic level fish
Secondary Trophic Transfer from
macroinvertebrates/
icthyoplankton/ other zooplankton
Initial Trophic Transfer
from phytoplankton, periphyton, macrophytes, detritus, & sediment
Selenium in Water Column
Selenium Sources
Naturally elevated selenium in soils - agricultural irrigation practices (Western US only)
Mining activities - coal, metals and sulfide minerals, phosphate
Algal/Plant Transformation and Enrichment:
 As function of sorption to particulates (sediment, algae, detritus)
As function of system hydrodynamics, lotic & lentic systems, residence time


	I	
Population decline
Figure 2.2. Diagram of Selenium Partitioning, Bioaccumulation, and Effects in the Aquatic
Environment.
The conceptual model links sources, transformation and uptake through media phases,
and consumer transfer and dynamics reflective of the movement of selenium through ecosystems
(Figure 2.2). Diet is the dominant pathway of selenium exposure for both invertebrates and
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vertebrates. Selenium moves from water to particulates, a collection of biotic and abiotic
compartments that includes primary producers, detritus, and sediments, which form the base of
aquatic food webs. Transfer from particulates to primary consumers (e.g., macroinvertebrates) to
fish is species specific. Knowledge of the food web is one of the keys to determining which
biological species or other ecological characteristics will be affected.
During the development of CWA section 304(a) criteria, EPA assembles all available test
data and considers all the relevant data that meet acceptable data quality and test acceptability
standards. This criterion update document is specific to selenium in fresh water. Chronic
criterion elements for selenium are protective concentrations measured in fish tissue and related
to protective water concentrations generated using food-web modeling. Further modeling is used
to estimate short-term concentrations in water from intermittent or pulsed exposures that are
protective against the chronic effect.
2.7.5 Analysis Plan for Derivation of the Chronic Fish Tissue-Based Criterion Elements
Data for possible inclusion in the selenium dataset were obtained primarily by search of
published literature using EPA's public ECOTOX database (up to July 2013). These studies were
screened for data quality as described in the EPA Ambient Water Quality Criteria Guidelines,
and adjusted for factors related to dietary lab or field exposure, which were not considered at the
time the Guidelines were written. Additional data were considered and reviewed for inclusion in
this criterion based on the public and peer review comments on the 2014 "External Peer Review
Draft" criterion document, and public comments on the 2015 draft.
Chronic toxicity studies (both laboratory and field studies) were further screened to
ensure they contained the relevant chronic exposure conditions of selenium to aquatic organisms
(i.e., dietary, or dietary and waterborne selenium exposure), measurement of chronic effects, and
measurement of selenium in tissue(s). The criterion derivation uses only those studies in which
test organisms were exposed to selenium in their diet, because such studies most closely replicate
real-world exposures (diet and/or diet plus water). This approach accords with findings and
recommendations of the 2009 SETAC Pellston Workshop (Chapman et al. 2009, 2010).
EPA grouped studies based on whether the effects were chronic reproductive (e.g.,
effects on offspring survival or morphology) or chronic non-reproductive (e.g., juvenile growth
and survival). At the 2009 Pellston workshop (Chapman et al. 2009, 2010), a group of 46 experts
in the area of ecological assessment of selenium in the aquatic environment agreed that the most
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important toxicological effects of selenium in fish arise following maternal transfer of selenium
to eggs during vitellogenesis, resulting in selenium exposure when hatched larvae undergo yolk
absorption. Such effects include larval mortality or permanent developmental malformations,
such as skeletal and craniofacial deformities. Therefore, the chronic fish-tissue-based criterion
elements are based on reproductive effects only.
The egg-ovary Species Mean Chronic Values (SMCVs) were calculated from the chronic
values (ECios and occasionally NOECs) obtained from the relevant toxicity tests. Genus Mean
Chronic Values (GMCVs) were calculated from the SMCVs and then rank-ordered from least to
most sensitive. The four lowest egg-ovary Genus Mean Chronic Values (GMCVs), whose exact
values influence the calculation of the egg-ovary criterion element, are all based solely on ECios.
The egg-ovary Final Chronic Value (FCV) was calculated from regression analysis of the four
most sensitive GMCVs, in this case extrapolating to the 5th percentile of the distribution
represented by the tested genera. The FCV directly serves as the fish tissue egg-ovary criterion
concentration element without further adjustment because the underlying ECios represent a low
level of effect (per the EPA Ambient Water Quality Criteria Guidelines).
For the whole-body and muscle criterion element concentrations, CVs were either
measured directly using the relevant tissue or the egg-ovary CVs were converted to estimated
equivalent whole-body or muscle CVs. The criterion concentration element expressed as whole-
body or as muscle concentration was calculated in a manner similar to the egg-ovary criterion
element using a combination of directly calculated CVs or CVs that used conversion factors
described below.
2.7.6 Analysis Plan for Derivation of the Fish Tissue Criterion Elements Duration
Duration of the averaging periods in national criteria restrict allowable fluctuations in the
concentration of the pollutant in the receiving water and restrict the length of time that the
concentration in the receiving water can be continuously above a criterion concentration, in order
to protect aquatic life. A numerical value for the fish tissue criterion elements averaging period,
or duration, is specified as instantaneous, because fish tissue data provide point, or instantaneous,
measurements that reflect integrative accumulation of selenium over time and space in the fish
population(s) at a given site. Selenium concentrations in fish tissue are generally expected to
change only gradually over time (Section 3.2.6 and Appendix J) in response to environmental
fluctuations; thus, there would be relatively little difference in tissue concentrations with
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different averaging period durations if the average selenium concentrations in water are
relatively stable over time. Generally fish collected to calculate average tissue concentrations for
a site are collected in one sampling event, or over a short time interval due to logistical
constraints and costs for obtaining samples incurred by state monitoring programs.
2.7.7 Analysis Plan for Derivation of the Fish Tissue Criterion Elements Return Frequency
Frequency is the number of times an excursion can occur over time without impairing the
aquatic community or other use. The current recommendation (1985 Guidelines - EPA PB85-
227049) for return frequency of once in 3 years on average is based on the ability of an aquatic
ecosystem to recover from a toxic insult when pollutant impacts are associated exclusively with a
water column exposure. This recommendation is based on the variability of water concentrations
that aquatic life will be exposed to, and is set at a low level such that the water concentrations
would mostly be below the criteria concentration. Selenium, however, is a bioaccumulative
pollutant, and elevated levels in various ecological compartments (e.g., biota, surficial
sediments) require a long period to decrease and the associated aquatic community requires a
long time to recover following reduction or removal of an elevated selenium exposure to a given
system (e.g., Belews Lake, NC, and Hyco Lake, NC).
Cumbie and Van Horn (1978) first reported young of the year losses in Belews Lake
quickly followed by dramatic decreases in standing stocks of many species, and particularly
game species like bluegill and largemouth bass. Fish communities were reduced to selenium-
tolerant species including cyprinids (e.g., fathead minnow) and green sunfish in both lakes.
Selenium reduction in Belews Lake (1985) and Hyco Lake (1990), resulted in rapid decreases in
[Se] in the water column, but reductions in fish tissue took much longer. Finley and Garrett
(2007) show that concentrations in bluegill and largemouth decreased from 19 and 17 mg/kg dw,
respectively in 1992-1994 to -8.0 mg/kg dw in both species sampled between 2003-2005. In
Belews Lake, where Se contamination was higher, [Se] in crappie and redear sunfish averaged
18 and 17 mg/kg dw respectively in 1994-1996, and decreased to -9-10 mg/kg dw in both
species based on sampling in 2004-2006, twenty years later.
Chapman et al. (2010) also reported a similar scenario for Hyco Lake where "fish
recruitment failure and the a massive fish kill in 1980 led to a decimated fishery. Selenium
concentrations in the reservoir were reduced beginning in 1990 and gradual reductions in Se
exposure via the food web led to the reestablishment of a diverse Hyco Lake fish community
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similar to the period prior to Se impact." The Belews and Hyco Lake examples indicate that a
protracted period of time (in excess of 10 years) would be necessary for fish communities to
recover once a selenium in fish tissue reached concentrations associated with reproductive
impacts. Thus, the typical "once-in-three years on average" criteria return frequency is not
appropriate for selenium, as this could lead to sustained ecological impacts.
2.7.8 Analysis Plan for Derivation of Chronic Water-based Criterion Element
The relationship between the ambient concentration of selenium in water and the
concentration of selenium in the eggs or ovaries of fish is primarily through trophic transfer of
selenium, which is greatly affected by site-specific conditions. EPA used a peer-reviewed model
to derive water concentrations from the egg-ovary criterion element that explicitly recognizes
partitioning of selenium in water and particulate material (algae, detritus, and sediment), and
trophic transfer from particulate material to aquatic invertebrates, from invertebrates to fish, and
partitioning in fish whole-body and fish eggs and ovaries. The method is composed of five main
steps:
1.	Formulate a mathematical equation relating the concentration of selenium in the eggs and
ovaries of fish to the ambient concentration of selenium in the water column.
2.	Develop parameters needed to use the mathematical equation formulated in step 1 from
available empirical or laboratory data related to selenium bioaccumulation in aquatic systems
and aquatic organisms.
3.	Classify categories of aquatic systems where a single water column concentration would be
adequately protective by evaluating the bioaccumulation potential at the base of the aquatic
food web.
4.	Translate the egg-ovary criterion element to an equivalent water column concentration at
each aquatic site.
5.	Apply a statistical threshold to the distribution of translated water column concentrations for
each aquatic system category to yield a water column concentration value that would be
protective of each aquatic system category.
EPA worked with USGS to derive a translation equation to estimate the site-specific
concentration of selenium in the water column corresponding to the egg-ovary criterion element
concentration. This equation utilizes a mechanistic model of bioaccumulation previously
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published in peer-reviewed scientific literature (Luoma et. al. 1992; Wang et. al. 1996; Luoma
and Fisher 1997; Wang 2001; Schlekat et al. 2002b; Luoma and Rainbow 2005; Presser and
Luoma 2006, 2010; Presser 2013). The equation uses site-specific food web models, species-
specific Trophic Transfer Factor (TTF) values, egg-ovary to whole-body conversion factor (CF)
values, and a site-specific enrichment factor (EF) values to calculate a site-specific water column
concentration element from the egg-ovary criterion element.
Empirical or laboratory data related to selenium bioaccumulation in aquatic organisms
are needed to calculate species-specific TTF and CF parameters and a site-specific EF parameter.
EPA obtained these data by reviewing its extensive selenium library of published papers and
reports, by searching published literature using EPA's public ECOTOX database and other
publically available data received through solicitation of public comments on the 2014 "External
Peer Review" draft, through the external peer reviewers of the 2014 draft, and through public
comments on the 2015 draft criterion document. Studies were screened using the same data
quality guidelines described above. Relevant studies contained selenium measurements from
field studies (water, particulate material, and aquatic organisms) or contained laboratory data on
physiological parameters of selenium bioaccumulation in aquatic organisms. Literature searches
for information on selenium associated with particulate matter included searches for data on all
forms of algae, detritus, inorganic suspended material, and sediment.
EPA compiled a collection of selenium concentration measurements from acceptable
field studies. Measurements were accepted if the study indicated the samples were collected in
the field, and the study identified the unit of measure, the media from which the measurement
was made, the location from where the sample was taken, and the date the sample was collected.
EPA only used data from studies with adequately described field collection protocols and where
concentrations were within the bounds of concentrations found using modern, rigorous protocols
in similar systems (Sanudo-Wilhelmy et al. 2004). The spatial precision of field data sample
collection locations were generally at the site level, although aggregate measurements were also
included if exposure conditions were considered similar (e.g., averages of single or composite
measurements from several locations in the same aquatic system). The temporal precision of
sample collection times were usually at the level of the day they were collected, although some
studies only provided enough information to determine the week, month, or year. If the day a
series of samples were collected was not reported but the study provided information that
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indicated the samples were taken concurrently, EPA noted sample precision, but assigned a
single effective collection date to all the samples.
EPA also compiled a collection of physiological coefficients for food ingestion rate (IR),
selenium assimilation efficiency (A A'), and rate of selenium loss (ke) from published literature.
Coefficients were accepted if the studies provided either the actual measurements or sufficient
information to derive them, and were reported in standard units (ke: /d; AE: %; IR: g/g-d) or
could be converted to standard units. Even though IR can be highly variable (Whitledge and
Hayward 2000), IR values of surrogate species were occasionally used.
EPA accounted for bioaccumulation variability across aquatic sites by evaluating the
parameter EF (representing the partitioning of selenium between the dissolved and particulate
state) from representative aquatic systems. The parameter EF is a measure of bioaccumulation
potential because it quantifies the transfer of selenium from the water column to particulate
material, which is the single most influential step in selenium bioaccumulation (Chapman et al.
2010). EPA calculated EF values for a set of aquatic systems using data from published literature
and applied statistical methods to distinguish categories with similar bioaccumulation
characteristics. On this basis, a single water column concentration is deemed adequately
protective when it is derived using data from aquatic sites in the same category. EPA translated
the egg-ovary criterion element into a set of water concentration values and derived a water
column criterion element for each aquatic system category using a percentile of the water column
concentrations for each category. To ensure adequate protection, EPA selected the 20th percentile
of the distribution of median water column values as the statistical cut-off (see Section 3.2.5),
Figure 2.3 diagrams the conceptual framework EPA used to derive water column criterion
element values from the egg-ovary criterion element.
2.7.9 Analysis Plan for Derivation of the Water Criterion Elements Duration
A numerical value for the lentic and lotic water criterion elements averaging period, or
duration, is specified as a 30-day average, because the presence of selenium in water is the initial
step in the process of bioaccumulation from the water column to fish tissue. The
bioaccumulation process for selenium takes place over a longer term than typically observed for
acute and chronic effects on aquatic life based on water concentrations. The derivation of a
protective water averaging period from kinetic modeling considerations is described in Section
3.2.6 and in Appendix J. Because the intermittent criterion element values are based on the water
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criteria chronic magnitudes and duration, the kinetic analysis of Appendix J also controls the
intermittent criterion element values.
Egg-ovary criterion element

Representative aquatic systems
Egg/ovary - whole body Conversion Factors (CP)
Concentrations in whole-body of fish
Trophic Transfer Functions (TIF)
O
Concentrations in prey
Trophic Transfer Functions (277)
O
Concentrations in particulate material
Site-specific Enrichment Factors (EF)
O
Distribution of water concentrations
Statistical threshold
O
Water criterion element
Figure 2.3. Conceptual Model for Translating the Selenium Egg-Ovary Concentration to a
Water Column Concentration.
2.7.10 Analysis Plan for Intermittent-Exposure Water-based Criterion Element Derivation
Like the chronic water criterion element, the intermittent-exposure criterion element
protects against cumulative exposure of selenium from multiple short-term discharges that may
cause an excursion of the fish tissue criterion element. EPA derived the intermittent exposure
criterion element directly from the chronic water criterion element by algebraically rearranging
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the chronic water criterion element to establish a limit on an intermittent elevated concentration
occurring over a specified percentage of time, while simultaneously accounting for background
concentrations (see Section 3.3).
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3 Effects Analysis for Freshwater Aquatic Organisms
3.1 Chronic Tissue-Based Selenium Criterion Element Concentration
Data were obtained primarily by search of published literature using EPA's public
ECOTOX database. The most recent ECOTOX database search extended to July 2013; this
document also reflects data either gathered or received by EPA based on information from the
2014	public comment period and 2014 external expert peer review of the "External Peer Review
Draft" published in May 2014, as well as information gathered based on public comments on the
2015	draft criterion. All available, relevant, and reliable chronic toxicity values were
incorporated into the appropriate selenium AWQC tables and used to recalculate the FCV, as
outlined in detail in the EPA Ambient Water Quality Criteria Guidelines.
The chronic values determined from acceptable chronic toxicity studies were separated
into reproductive endpoint and non-reproductive endpoint categories. Although both sets of
endpoints assess effects due to selenium on embryo/larval or juvenile development and survival
and growth, the fundamental difference is exposure route (inherent in test design). That is, the
fundamental difference is whether the aquatic organisms (e.g., fish) were directly exposed to
selenium in the diet and water column or exposed via maternal transfer of selenium to the
eggs/ovaries prior to reproduction. In studies with reproductive endpoints, parental females are
exposed to selenium and the contaminant is transferred from the female to her eggs. In the
selenium-exposed females, selenium replaces sulfur in vitellogenin, the primary yolk precursor,
which is transported to the ovary and incorporated into the developing ovarian follicle (Janz et al.
2010). In most but not all of these studies, progeny from these females were not additionally
exposed to aqueous selenium. The chronic values derived for the reproductive effects (survival,
deformities, and edema) are based on the concentration of selenium in the eggs or ovary, the
tissues most directly associated with the observed effects. In contrast, in studies grouped under
non-reproductive effects (usually larval and/or juvenile survival or growth), the tested fish had
no maternal pre-exposure to selenium. Chronic values for non-reproductive effects are based on
the concentration of selenium in tissues measured in the study: muscle, liver and/or whole body.
The reproductive endpoint studies applied to the derivation of the chronic criterion
elements are described below. Less definitive reproductive studies that are not directly applied to
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the criterion derivation are described in Section 6.1.2 and in Appendix C. Nonreproductive
studies are described in Section 6.1.9,
3.1.1 Acceptable Studies of Fish Reproductive Effects for the Four Most Sensitive Genera
Below is a brief synopsis of the experimental design, test duration, relevant test
endpoints, and other critical information regarding the four sensitive genera that drive the
calculation of each specific chronic value. The studies in this section involve effects on the
offspring of exposed female fish. Data are summarized in Table 3.1. Details of these studies and
other chronic studies considered for criterion derivation are contained in Appendix C.
3.1.1.1 Acipenseridae
Acipenser transmontanus (white sturgeon)
Linville (2006) evaluated the effect of elevated dietary selenium on the health and
reproduction of white sturgeon. Adult female white sturgeon (approximately 5 years old) were
fed either a control diet (no added selenium, 1.4 mg/kg Se) or a diet spiked with selenized yeast
(34 mg/kg Se) for six months in a freshwater flow through system. At the end of the dietary
exposure, females were induced to spawn and fertilized with non-exposed male milt. Large
cohorts of fertilized eggs from individual females (two from control and three from the
treatment) were collected and separately hatched. After hatching (stage 36), n=500 sets of larvae
were randomly distributed into each of six flowthrough chambers, three for stage 40 assessment
and three for stage 45 assessment. Length, weight, mortality, abnormalities (edema, skeletal
deformities) and selenium were measured at stages 36, 40 and 45. The mortality and abnormality
observations from oldest stage (45) were used for effects analysis because these measurements
showed the greatest response.
No selenium effects were observed for length or weight of larvae but effects were
observed for both abnormalities (edema and skeletal deformities) and survival. Selenium
concentrations in eggs from the control fish were 1.61 and 2.68 mg/kg dry weight (dw), and were
7.61, 11 and 20.5 mg/kg dw in eggs from the treatment fish. As stated above larval survival and
abnormality frequency was evaluated at stage 45. Because the mortalities for each cohort were
recorded up to the sample collection time for abnormalities, a combined effects variable was
derived based on the total proportion of hatched larvae which were both alive and without any
abnormalities at stage 45. This can be calculated as PS*(1-PA), where PS is the proportion
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survival in the test chambers and PA is the proportion of the sample of surviving larvae with
abnormalities. The larvae hatched from the batches of eggs with selenium concentrations of 1.61,
2.68,7.61, 11 and 20.5 mg Se/kg dw had 0.3, 0.3, 13.6,0.3 and 33.8% combined survival and
abnormal (edema and deformities) effects, respectively.
Estimation of the ECio was conducted using weighted nonlinear regression analysis with
the threshold sigmoid model equation (TRAP version 1.30a). The binary data (i.e., survival and
abnormalities) available in this study would normally be analyzed using the tolerance
distribution analysis in TRAP; however, the combined survival/abnormalities effects variable in
this study precludes its use because of the different sample sizes for survival and abnormality
evaluation. When there are insufficient data for TRAP to fit an effects/exposure curve, an
interpolation is conducted with the same TRAP equation, but constrained to provide
interpolation between two points.
Since the study yielded only one definite partial effect, TRAP cannot be used to estimate
a concentration-response curve. Instead, TRAP was used to interpolate between the last two
points to estimate the ECio (see Linville summary in Appendix C for detail). The resultant TRAP
slope is 2.96 and the interpolated ECio is 15.6 mg/kg.
The white sturgeon ECio of 15.6 mg/kg egg dw is important to include in the criterion
analyses because this species a commercially and recreationally important fish species in the
Pacific Northwest, and also serves as a surrogate for other sturgeon species in the United States
(see Section 6.4, Protection of Threatened or Endangered Species), and has a population listed as
endangered in the Kootenai River in Idaho and Montana.
3.1.1.2 Salmonidae
Acceptable studies were available for three salmonid genera, Oncorhynchus, Salvelinus
and Salmo. All of these studies evaluated the effects of selenium on salmonid embryo/larval
survival and deformity and used wild-caught adults taken from selenium contaminated streams
and spawned for effects determination. Exposure for all studies was therefore through the
parents. Summaries of the studies with Salvelinus are discussed in Section 6.1.2.3;
Oncorhynchus and brown trout {Salmo trutta) are discussed below.
Oncorhynchus mykiss (rainbow trout) Holm (2002) and Holm et al. (2005) obtained eggs
and milt from ripe rainbow trout collected from reference streams and streams containing
elevated selenium from an active coal mine in Alberta, Canada. In 2000, 2001 and 2002 eggs
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were fertilized and monitored in the laboratory until swim-up stage, for percent fertilization,
deformities (craniofacial, finfold, and spinal malformations), edema, and mortality. No
significant differences among sites were observed for percent fertilization and mortality.
Percentages of embryonic deformities and edema were significantly different among streams, but
rates of deformities at Wampus Creek, one of the reference streams, were often similar to or
higher than deformities at streams with elevated concentrations of selenium (see Holm summary
in Appendix C). The measurement of selenium in the otolith layers of rainbow trout collected in
this watershed showed low selenium exposure in the fish's early life and a higher exposure to
selenium during the fish's adult years (Palace et al. 2007), suggesting that individuals that reach
adulthood tend not to start their lives in streams with elevated selenium concentrations, even
though they may reside there later.
Craniofacial deformities, skeletal deformities and edema in rainbow trout embryo, as a
function of selenium in egg wet weight (ww), were fitted to a curve using a weighted regression
and a threshold sigmoidal equation (TRAP) from which ECio values were calculated (see
Appendix C for tables and figures). EC estimates for finfold deformities, length and weight of
rainbow trout embryos could not be made because of inadequate dose-response data. The most
sensitive endpoint was edema with an ECio value of 9.5 mg Se/kg egg ww or 24.5 mg Se/kg egg
dw. The conversion of ww to dw used the average percent moisture of 61.2% for rainbow trout
eggs reported by Seilor and Skorupa (2001).
Oncorhynchus clarkii lewisi (Westslope cutthroat trout)
In a field study similar to those conducted by Holm et al. (2005), Rudolph et al. (2008)
collected eggs from Westslope cutthroat trout from Clode Pond (exposed site) and O'Rourke
Lake (reference site). Clode Pond is on the property of Fording River Coal Operations in
Southeast British Columbia with reported selenium concentrations of 93 |ig/L. O'Rourke Lake is
an isolated water body into which Westslope cutthroat trout were stocked in 1985, 1989 and
1992 and has selenium levels reported as <1 |ig/L. Eggs with the four highest Se concentrations
(86.3 to 140 mg/kg dw) collected from Clode Pond fish died before reaching the laboratory. Of
those eggs from both ponds that survived, there was no correlation between egg selenium
concentration and frequency of deformity or edema in the fry. The percent of alevins (post hatch
to swim-up stage) that died was related to the selenium concentration in the eggs; the ECio
estimate for alevin survival based on the concentration of selenium in the eggs is 24.7 mg Se/kg
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dw. See Appendix C for details on the statistical analysis and how it differed from the previous
draft(s).
As a follow-up to the study by Rudolph et al. (2008), Nautilus Environmental (2011)
conducted a more extensive study with Westslope cutthroat trout at the same site. Adult
Westslope cutthroat trout were collected from lentic and lotic environments from locations near
the mining operations. The lentic fish were primarily captured in Clode Pond, a settling area used
to improve water quality of the mining discharge. Lotic fish were collected from the Fording
River and its tributaries near the mining operation. Reference females were obtained from
Connor Lake which is located within the watershed but not exposed to mining discharges. The
researchers reared fertilized eggs from the caught females in the laboratory until they reached
swim-up fry stage. A subset of fry surviving at swim-up were reared for an additional 28 days.
The most sensitive endpoint was larval survival at swim-up with an ECio of 27.7 mg/kg egg dw
determined by interpolation between the one partial effect (20.8% survival at 34.2 mg Se/kg and
average NOEC of 87.2% survival at 24.8 mg Se/kg; see Appendix C for detail and how this
statistical analysis differed from the previous draft(s)). This result is very similar to the ECio of
24.7 mg/kg egg dw determined for the data generated by Rudolph et al. (2008). See Appendix C
for more details on the Nautilus Environmental (2011) study.
The GMCV for Oncorhynchus reproductive endpoints is 25.3 mg Se/kg EO. This GMCV
is the geometric mean of the O. mykiss ECio of 24.5 mg (Holm 2002 and Holm et al. 2005) and
the SMCV of 26.2 mg Se/kg EO dw for 0. clarkii. The 0. clarkii SMCV was based on the ECio
values of 24.7 mg Se/kg EO from Rudolph et al (2008) and 27.7 mg Se/kg EO dw from Nautilus
Environmental (2011).
Salmo trutta (brown trout)
Formation Environmental (2011) collected adult female and male brown trout from sites
with low and high selenium exposure in the vicinity of a phosphate mine located in Southeastern
Idaho in November 2007. Eggs were collected from 26 gravid females across three sampling
locations, fertilized with milt collected from several males from the same site and taken to the
laboratory for hatching and observation of larval malformations and survival. In addition to the
field collected fish, fertilized eggs of 14 females from two separate hatcheries were used in the
study.
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The study had two phases, hatch-to-swim up, and swim up-to-15 days post swim-up.
There are two experimental complications that affect the interpretation of these data: (a) elevated
deformity rates among the offspring that were to serve as hatchery-originated method controls
(very low selenium exposure) and among some of the low exposure field-collected organisms,
and (b) the accidental loss of a number of individuals from several treatments during the 15-day
post swim up portion of the test due to overflow of the tank water. The current criterion
document's analysis is still based on the revised count data from AECOM (2012), which built
upon and superseded EPA's 2012 analysis (Taulbee et al. 2012), peer reviewed by ERG (2012).
Approximately 600 eggs/female were obtained from the majority of the field and
hatchery-collected females; however, the numbers of eggs per female ranged from 20 to 609.
Selenium concentrations were measured for a subsample of eggs taken from each field and
hatchery-collected female. EPA's primary evaluation in this document is of the survival of larvae
from hatch to swim-up. Hatching success and larval survival were monitored to swim-up, at
which time the fry were thinned to a maximum of 100 individuals for monitoring survival for 15
days post swim-up. Larvae from 24 field collected and 14 hatchery collected females were
assessed for survival, as no larvae hatched from the eggs of two of the 26 field collected females.
Because of uncertainties regarding how best to address the loss of fish during the
overflow event during the second phase of the test, and also because of the preferential selection
of healthy fish during the thinning process prior to the post-swim-up portion of the test, where
only those individuals presumed to be healthy were retained for assessment of deformities, EPA
used survival during only the first portion of the test (hatch to swim-up), as it provides a more
reliable chronic value.
The dataset of percent survival from hatch to swim-up versus the selenium concentration
in eggs is an excellent dataset and provides a good foundation for setting numeric effect
concentrations for selenium. There is a narrow range between the NOEC (20.5 mg/kg) and a
LOEC with severe effects (26.8 mg/kg, 61% mortality) that leaves little uncertainty in what an
appropriately protective effects concentration should be. There are sufficient data for TRAP to
estimate a curve, using weighted least-squares nonlinear regression with the threshold sigmoidal
model. The weighting factor for the 33 no-effect points is their standard deviation, and the
weighting factor for the 5 effect points is their residual standard deviation. The TRAP parameter
values are 96.2% for background survival, 1.45 for the logECso (ECso=28.2 mg/kg), and 4.28 for
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the slope. The ECio is estimated to be 21.0 mg/kg, slightly higher than the NOEC of 20.5 mg/kg.
The weighted TRAP model curve fits the 5 higher effects data well, which forces the ECo
estimate down to 16.4 mg/kg, below two of the points in the background range. In particular, the
fitted curve goes through the NOEC data point at 20.5 mg/kg, so that this point is considered to
be an ECg. This is reasonable because the response is so steep at concentrations above this point
that some effect at this point is plausible, and also provides further support of an ECio at 21.0
mg/kg. Section 6.1.6 provides a summary of the analysis that led to the final selection of the
ECio for larval survival during the first portion of the test. Appendix C presents details of the
study and analysis.
3.1.1.3 C entrarchi dae
Lepomis macrochirus (bluegill sunfish)
In a laboratory study, Doroshov et al. (1992a) exposed adult bluegill for 140 days to three
dietary treatments of seleno-L-methionine (nominal dietary concentrations of 8, 18 and 28 mg
Se/kg) added to trout chow. Near the end of the exposure, ripe females were induced to ovulate
and ova were fertilized in vitro with milt stripped from males. Fertilized eggs were sampled for
fertilization success and selenium content. They were also used in two tests, (a) a larval
development study during the first 5 days after hatching, and (b) a 30-day embryo-larval test. In
the 30-day larval survival test, statistical difference from the control was only found in the
highest test treatment for survival and growth (length and weight) measurements. In the 5-day
larval test, the average proportion of larvae with edema was 0% at an egg concentration of 8.33
mg Se/kg (8 mg/kg dietary treatment), 5% at an egg concentration of 19.46 mg Se/kg dw (18
mg/kg dietary treatment), and 95% at an egg concentration of 38.39 mg Se/kg dw (28 mg/kg
dietary treatment). The latter two were statistically different from the control (0% edema). All
edematous larvae died in the high treatment.
This analysis focuses on the available data for the individual replicates for the 4-day data
(5-day were not used because of almost complete mortality at the highest treatment). Of the 33
edema measurements, only 15 could be used because not all the replicate egg concentrations
were reported. Table 4 in the Doroshov et al. (1992a) summary in Appendix C shows both
individual replicates and the treatment averages, which are only slightly different than the 5-day
data (averages) previously used in the selenium document. Individual replicates rather than
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treatment means were used because the exposure concentrations vary substantially and effects
are correlated with exposure within the treatment (illustrated by nominal dietary treatments of 18
mg/kg (with corresponding Se concentrations in eggs at that nominal treatment level ranging
from 8.55 to 30.20 mg/kg) and 28 mg/kg (with corresponding Se concentrations in eggs ranging
from 25.21 to 52.18 mg/kg; see Appendix C for details).
TRAP was fitted to the available data based on the individual replicates and the treatment
means using the tolerance distribution option with the log-triangular distribution. In both cases,
the TRAP program indicates that the dataset contains inadequate partial responses because the
partial responses are less than 10% or greater than 90%, and there are no data (responses)
between 10 and 90%. However, for this dataset, these partial responses at both ends of the
concentration response curve are sufficiently informative based on multiple lines of evidence
(e.g., same response on both days 4 and 5, other endpoints that show effects at dietary treatment
of 18 mg/kg, and several instances of edema at dietary treatment 18 mg/kg in contrast to
absolutely none for many observations at any lower concentration). Also, because dietary
treatment 18 mg/kg does have an effect of several percent or so, estimating the ECio near these
points is defensible. Therefore, the ECio estimated using separate replicates is 22.6 mg/kg dw.
A similar study with similar results was done by Coyle et al. (1993) in which two year
old pond-reared bluegill sunfish were exposed in the laboratory and fed (twice daily ad libitum)
Oregon moist pellets containing increasing concentrations of seleno-L-methionine. Water
concentrations were nominal 10 |ig Se/L. The fish were grown under these test conditions for
140 days. Spawning frequency, fecundity, and percentage hatch were monitored for 60 days
from the initiation of spawning. There was no effect from the highest dietary selenium
concentration (33.3 mg Se/kg dw) on adult growth, condition factor, gonadal somatic index, or
other endpoints (Appendix C). The effect of interest in this study was 30-d larval survival after
hatch (deformities weren't examined and other reproductive endpoints showed no effect at the
highest exposure). For this survival endpoint, there was complete mortality after one week at the
highest exposure and no significant differences in survival at lower concentrations.
Previously, the day 5 data were used in the analysis described in Appendix C because this
was the only day in which control survival was greater than 90%, with the control and all the
treatments showing substantial and increasing toxicity over the next 4 days. However, upon
closer analysis, EPA asserts that this is not sufficient cause to base the assessment, because from
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day 6 through day 30, survival at the fifth treatment was greater than survival in the first and
third treatments, indicating this is not an effect level. These later data (day 6-30) establish that
the highest treatment is best considered an ECioo and the fifth treatment an ECo. Therefore, the
TRAP interpolation was redone using 42 mg/kg as an ECioo rather than an EC93, resulting in a
slope of 7.6 and an EC10 of 26.3 mg Se/kg dw in eggs.
Hermanutz et al. (1992), and Hermanutz et al. (1996) exposed bluegill sunfish to sodium
selenite spiked into artificial streams (nominal test concentrations: 0, 2.5, 10, and 30 |ig Se/L)
which entered the food web, thus providing a simulated field exposure (waterborne and dietary
selenium exposure). In an effort originally intended to improve the rigor of the statistical analysis
of the Hermanutz et al. (1996) data, Tao et al. (1999) re-examined the raw data records and made
corrections to the counts. This criterion document considers the Hermanutz et al. (1992) data and
the Tao et al. (1999) re-examination of Hermanutz et al. (1996).
These data come from a series of three studies lasting from 8 to 11 months, conducted
over a 3-year period. All three studies began with exposure of adult bluegill sunfish in the fall,
and with respective studies ending in the summer of the following year. Temperatures averaged
4.6, 4.1 and 4.5C during the winter months, and averaged 26.4, 23.9 and 22.4C during the
spawning months (June-July) for Studies I, II and III, respectively. Spawning activity was
monitored in the stream, and embryo and larval observations were made in situ and from
fertilized eggs taken from the streams and incubated within egg cups in the laboratory. None of
the adult bluegill exposed to the highest concentration of selenium in the water (Study I, mean
measured concentration equal to 29.4 |ig/L) survived the entire exposure period (although a few
did survive to spawn). Reduced survival and increased deformities on offspring were observed in
the selenium-dosed streams in both Study I and Study II, but were not found during Study III
(recovering from contamination, no active selenium input/treatment). The incidence of edema,
lordosis, hemorrhage and larval survival in the one stream concentration common to both Study I
and II, 10 |ig/L, ranged from 80 to 100 percent, 5 to 18 percent, 27 to 56 percent, and 29 to 58
percent, respectively over the three years (combined egg cup and nest observations). Edema,
lordosis, and hemorrhage in the lowest stream concentration in Study II, 2.5 |ig/L, ranged from 0
to 4 percent, 0 to 25 percent, and 3.6 to 75 percent, respectively (combined egg cup and nest
observations); larval survival was 71.6 percent (72 and 75 percent in the control streams). (See
Hermanutz 1996 and 1992 in Appendix C for more detail). The effects were not observed in
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larvae from fish that were not exposed to elevated concentrations of selenium (control
treatment). The mean concentrations of selenium in bluegill ovaries, measured at the end of each
study, ranged from 2.2 to 5.0 mg/kg dw in controls, 7.6 to 14 mg/kg dw in the 2.5 |ig/L
treatments, 34 to 52 mg/kg dw in the 10 |ig Se/L treatments, and 16 to 55 mg/kg dw in
recovering 30 |ig/L treatments. Muscle and whole-body measurements were also available. For
all three tissue types, concentrations measured during the spring of each exposure period were
not used because they were not sufficiently co-occurrent with the observation of effects. It
should also be noted that in contrast to more recent field studies, the tissue concentrations cannot
be linked from particular adult females to effects on her offspring, but only from an aliquot of
females in the treatment to all offspring observed in the treatment.
The egg-cup data for all streams of Studies I, II, and III of this experiment were
combined and analyzed in response to measured selenium concentration in the maternal ovaries
(mg/kg dw) using TRAP. That is, data for streams receiving water-borne selenium were
combined with data for streams recovering from the previous year's contamination. The absence
of effects at high tissue levels (55 mg Se/kg ovary dw) in the recovering stream of Study II did
not affect the ECio estimate because it was outweighed by three other points showing severe
effects at concentrations as low as 16.7 mg Se/kg ovary dw. However, this one observation is
suggestive but not definitive corroboration for the field observations of biological recovery in
Belews Lake and Hyco Reservoir after selenium loads were reduced, but while tissue
concentrations remained relatively high (Lemly 1997a; Crutchfield 2000; Finley and Garrett
2007).
Several egg-cup endpoints were analyzed by TRAP independently (% edema, % lordosis,
and % hemorrhage) and in combination (% normal and surviving). The best fit and most
sensitive was the combined percent normal and surviving larvae. Due to inadequate partial
effects for the ovary analysis, a threshold sigmoidal model was used to interpolate an ECio
estimate between the first interpolation point set to the highest no observed effect concentration
(HNOEC) of 14.0 mg/kg and the average background survival/normal of 69.1% and the second
point set to the LOEC of 16.7 mg/kg and a survival/normal of 5.76%. The resulting ECio is 14.7
mg/kg ovary dw. Chronic values for muscle and whole body based on percentage surviving and
free of deformities are 13.4 mg Se/kg muscle dw and 10.6 mg Se/kg whole body dw. (See
Appendix C for more discussion of this study).
43

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The SMCV for bluegill reproductive endpoints based on ECio values is 20.6 mg Se/kg
dw in egg/ovary, based on the ECio values of 22.6 mg/kg dw in the Doroshov et al. (1992a)
study, 26.3 mg/kg dw in the Coyle et al. (1993) study, and 14.7 mg/kg dw for Hermanutz et al.
(1992, and 1996 as corrected by Tao et al. 1999).
3.1.2 Summary of Acceptable Studies of Fish Reproductive Effects
Table 3.1 summarizes the effect concentrations obtained from all acceptable reproductive
studies with fish. Summaries of the remainder of the reproductive studies (beyond the four most
sensitive genera described above) can be found in Section 6.1.2 below.
44

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Table 3.1. Maternal Transfer Reproductive Toxicity Studies.
Species
Reference
Kxposurc Koule
Toxicological
Kndpoinl
Chronic Value
m/k d\v"
S.MCV
nig/k d\v
(i.MCV
nig/k d\v
Salvelinus malma
Dolly Varden
Golder 2009
dietary and
waterborne
(field: Kemess Mine
NW British Columbia)
EC io for total
deformities
56.2 E
56.2 E
56.2 E
Esox lucius
northern pike
Muscatello et
al. 2006
dietary and waterborne
(field: Saskatoon,
Sask.)
EC24 larval
deformities
34.0 E
34.0 E
34.0 E
Cyprinodon macularius
desert pupfish
Besser et al.
2012
dietary and waterborne
(lab)
Estimated EC 10 for
offspring survival
27 E
27 E
27 E
Micropterus salmoides
largemouth bass
Carolina Power
& Light 1997
dietary (lab)
EC10 for larval
mortality &
deformity
26.3 O
26.3 O
26.3 O
Pimephales promelas
fathead minnow
Schultz and
Hermanutz
1990
dietary and waterborne
(mesocosm:
Monticello)
LOEC for larval
edema and lordosis
<25.6 Eb
NA
NA
Oncorhynchus mykiss
rainbow trout
Holm 2002;
Holm et al.
2003, 2005
dietary and
waterborne
(field: Luscar River,
Alberta)
EC10 for skeletal
deformities
24.5 Eb
24.5 E
25.3 E
Oncorhynchus clarkii
lewisi
Westslope cutthroat trout
Rudolph et al.
2008
dietary and waterborne
(field: Clode Pond,
BC)
EC10 for alevin
mortality
24.7 E
26.2 E
Oncorhynchus clarkii
lewisi
Westslope cutthroat trout
Nautilus
Environmental
2011
dietary and waterborne
(field: Clode Pond &
Fording River, BC)
EC 10 for survival
at swim-up
27.7 E
Salmo trutta
brown trout
Formation
Environmental
2011; AECOM
2012
dietary and waterborne
(field: Lower Sage
Creek & Crow Creek,
ID)
EC10 for larval
survival
21.0 E
21.0 E
21.0 E
45

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Species
Reference
Kxposnre Uoulc
Toxicologicsil
Kml point
Chronic Value
m/k dw"
SMCV
nig/kg dw
CMCV
nij/k dw
Lepomis macrochirus
bluegill
Doroshov et al.
1992a
dietary (lab)
ECio larval edema
22.6 E
20.6 E
20.6 E
Lepomis macrochirus
bluegill
Coyle et al.
1993
dietary and waterborne
(lab)
ECio for larval
survival
26.3 E
Lepomis macrochirus
bluegill
Hermanutz et
al. 1992, 1996
dietary and waterborne
(mesocosm:
Monticello)
EC io for larval
edema
14.7 Ob
Acipenser transmontanus
white sturgeon
Linville 2006
dietary (lab)
ECio for combined
edema and
deformities
15.6 E
15.6E
15.6 E
E-Concentration reported in egg; O- concentration reported in ovary
a All chronic values reported in this table are based on the measured concentration of selenium in egg/ovary tissues.
b Tissue value converted from ww to dw. See Appendix C for conversion factors.
0 SMCV not calculated due to variability in the observations among replicates in Schultz and Hermanutz (1990). The chronic
value is presented in this table to show it is in the range of selenium effect concentrations. See Appendix C for detail. Also, see
Appendix E for an additional study with fathead minnow.
46

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In order of their sensitivity to selenium, Table 3.2 presents the Genus Mean Chronic
Values from acceptable fish reproductive-effect studies that have been measured in terms of egg-
ovary concentrations.
Table 3.2. Ranked Genus Mean Chronic Values for Fish Reproductive Effects Measured as
Egg or C
~vary Concentrations.
Knnk
(;m( \-
(mg Se/kg dw r.O)
Species
SMCV
(mg Se/kg dw KO)
8
56.2
Dolly Varden,
Salvelinus malma
56.2
7
24**
Northern pike,
Esox lucius
34
6
27
Desert pupfish,
Cyprinodon macularius
27
5
26.3
Largemouth bass,
Micropterus salmoides
26.3
4
25.3
Cutthroat trout,
Oncorhynchus clarkii
26.2
Rainbow trout,
Oncorhynchus mykiss
24.5
3
21.0
Brown trout,
Salmo trutta
21.0
2
20.6
Bluegill sunfish,
Lepomis macrochirus
20.6
1
15.6
White sturgeon,
Acipenser transmontanus
15.6
* This table excludes Gambusia, which has a reproductive chronic value expressed as adult
whole-body rather than egg-ovary, because it is a live bearer.
** The Northern Pike SMCV is an EC24 based on larval deformities (Table 3.1). The EC10 is less
than 34 mg/kg.
This table excludes GMCV for Pimephales due to uncertainty in the chronic value for the
Schultz and Hermanutz (1990) study; the estimate of <25.6 mg/kg egg dw in Table 3.1 shows it
is in the range of reproductive effect levels for selenium (See Appendix C for study details).
3.1.3 Derivation of Tissue Criterion Element Concentrations
Data used to derive the final chronic value were differentiated based on the effect
(reproductive and non-reproductive effects). Acceptable chronic toxicity data on fish
reproductive effects are available for 10 fish genera. Acceptable chronic toxicity data on non-
reproductive effects are available for 7 fish genera and 3 invertebrate genera. The fish non-
47

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reproductive effects data were not used to calculate tissue criterion elements because they were
more variable and less reproducible than the data on reproductive effects. The genus sensitivity
distribution is predominantly populated with data on fish species because field evidence
demonstrated that fish communities were affected in situations having no observable change in
the accompanying diverse array of invertebrate communities. As a result, decades of aquatic
toxicity research have focused primarily on fish. The studies that have been done with
invertebrates (Table 3.8, Section 3.1.4) have shown them to be more tolerant than most of the
tested fish species.
Also, while amphibians are potentially sensitive due to physiologic similarities to fish,
effects clearly attributable to selenium are largely unknown (Unrine et al. 2007; Hopkins et al.
2000; Janz et al. 2010). Hopkins et al. (2000) reported that amphibian larvae at sites receiving
coal combustion wastes appear to efficiently accumulate selenium in their tissues and possibly
due to selenium have exhibited axial malformations. In a recent laboratory exposure, Masse et al.
(2015) determined an ECio of 44.9 mg/kg Se for the African clawed frog (Xenopus laevis)
suggesting that amphibians are similar to the less sensitive fish species (see Section 6.1.4),
3.1.3.1 Fish Egg-Ovary Concentration
The lowest four GMCVs from Table 3.2 are shown below in Table 3.3.
Table 3.3. Four Lowest Genus Mean Chronic Values for Fish Reproductive Effects
Relative Sensitivity
Knnk
(ienus
cm (v
(m Se/kg dw e-ovarv)
4
Oncorhynchus
25.3
3
Salmo
21.0
2
Lepomis
20.6
1
Acipenser
15.6
With N=15 GMCVs (see Section 3.1.6), the 5th percentile projection yields an egg/ovary
criterion element concentration of 15.1 mg Se/kg dw egg/ovary, lower than the most sensitive
fish species tested, white sturgeon (A. transmontanus). The egg/ovary criterion element
concentration is compared to the distribution of egg/ovary chronic values in Figure 3.1.
48

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64
T3
two
 32
E
o
u
>
O
i
two
two

~
~
X
Salvelinus
Esox

X
X
Cyprinodon
+ - *

-
Micropterus
~ -

+
Oncorhynchus
Salmo
	

~

Lepomis
Acipenser
EO FCV
o ; n (>
8
10

Rank



Figure 3.1. Distribution of Reproductive-Effect GMCVs for Fish Measured as Egg-Ovary
Concentrations.
3.1.3.2 Fish Whole-Body Criterion Element Concentration
Whole body reproductive chronic values were calculated directly from whole body tissue
concentrations measured in the study or by applying an egg-ovary (EO) to whole-body (WB)
conversion factor (CF) presented subsequently in Section 3.2.2.2, Direct calculations were done
when whole body measurements were available in the study and the data were amenable to an
effect level determination. Table 3.4 provides the chronic values for each fish genus and whether
it was calculated directly or converted from the reproductive-effect egg-ovary concentrations to
whole-body concentrations using a CF. The final EO/WB CF applied to each taxon was
determined using a hierarchical approach based on taxonomic relatedness, and is described in
Section 3.2.2, and in greater detail in Appendix B. The four most sensitive reproductive-effect
fish whole-body GMCVs are shown in Table 3.5.
49

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Table 3.4. Tested Reproductive-Effect Whole Body (WB) Concentrations Measured
Directlv or Converted to WB Concentrations from Egg-Ovarv (EO) Concentrations.
Taxon-
r.o
Chronic
Value
KO/W'H
CI
Direct or
Calculated
W li Ucpro
Chronic
Value
Direct Calculation or
liasis for KOAYIi CI
(from Appendix 1$)
Salvelinus
56.2
1.61
34.9
Dolly Varden EO/M (1.26) x all fish
M/WB (1.27)
Esox
34.0
2.39
14.2
Northern pike EO/M (1.88) x all fish
M/WB (1.27)
Cyprinodon
27.0
1.20
22.6
Desert pupfish EO/WB
0. mykiss
24.5
2.44
10.0
Rainbow trout EO/M (1.92) x all
fish M/WB (1.27)
Rudolph et al.
2008
24.7
1.96
12.6
Oncorhynchus EO/WB
Nautilus 2011
27.7
1.96
14.1
Oncorhynchus EO/WB
O. clarkii
26.2
NA
13.3
Geometric mean of two studies
Oncorhynchus
25.3
NA
11.6
Geometric mean of O. mykiss and O.
clarkii WB SMCVs
Micropterus
26.3
1.42
18.5
Micropterus EO/WB
Salmo
21.0
NA
13.2
Directly calculated ECio
Coyleetal. 1993
26.3
NA
8.6
Directly calculated ECio
Doroshov et al.
1992a
22.6
2.13
10.6
Bluegill sunfish EO/WB
Hermanutz et al.
1992, 1996
14.7
NA
10.6
Directly calculated ECio
Lepomis
20.6
NA
9.9
Geometric mean of three studies
Acipenser
15.6
1.69
9.2
White sturgeon EO/M (1.33) x all
fish M/WB (1.27)
* The GMCV for Gambusia, a live bearer, not included in the conversion table, was originally
measured as adult WB, not EO, and is >13.38 mg Se/kg dw WB. The "greater than" sign
signifies that no effects were found at the highest observed concentrations. This table also
excludes Pimephales due to uncertainty in the chronic value for the Schultz and Hermanutz
(1990) study (See Appendix C for details).
50

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Table 3.5. The Lowest Four Reproductive-Effect Whole-Body GMCVs.
Relative
Sensitivity
Rank
Genus
GMCV
(mg Se/kg dw whole-body)
4
Salmo
13.2
3
Oncorhynchus
11.6
2
Lepomis
9.9
1
Acipenser
9.2
Because the factors used to convert egg-ovary to whole-body concentrations vary across
species, the whole-body rankings differ from the egg-ovary rankings. With N=15 GMCVs, the
5th percentile projection yields a whole body criterion element concentration of 8.5 mg Se/kg dw
whole-body, slightly lower than the most sensitive fish species tested, white sturgeon (Acipenser
transmontanus). The fish whole body criterion element is compared to the distribution of fish
whole-body reproductive chronic values in Figure 3.2.
~	Salvelinus
X	Cyprinodon
	Micropterus
X	Gambusia
+	Esox
Salmo
~	Oncorhynchus
A	Lepomis
	Acipenser
...	WBFCV
Rank
Figure 3.2. Distribution of Reproductive-Effect GMCVs for Fish, either Measured as
Whole-Body Concentrations in the Original Tests, or Measured as Egg-Ovary
Concentrations but Converted to Whole-Body.
(As shown in Table 3.4).

TS
32

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3.1.3.3 Fish Muscle Criterion Element Concentration
Reproductive chronic values for muscle tissue were calculated directly from muscle
tissue concentrations measured in the study or from the egg-ovary to muscle conversion factors
of the bioaccumulation modeling approach (presented in Section 3.2). Direct calculations were
made when muscle measurements were available in the study and the data were amenable to an
effect level determination. The final EO/M CF applied to each taxon was determined using a
hierarchical approach based on taxonomic relatedness, consistent with the approach used to
calculate EO/WB CFs described in Section 3.2.2,
Table 3.6 provides the chronic values for each fish taxa and whether it was calculated
directly or converted from reproductive-effect egg-ovary concentrations to muscle
concentrations. The four most sensitive reproductive-effect fish muscle GMCVs are shown in
Table 3.7
Table 3.6. Tested Reproductive-Effect Muscle (M) Concentrations Measured Directly or
Converted to M Concentrations from Egg-Ovary (EO) Concentrations.	
Taxon
i:o
Chronic
Value
EO/M
CI
Direct or
Calculated
Muscle Uepro
Chronic Value
Direct Calculation or
Basis for EO/M CI
(from Appendix 1$)
Salvelinus
56.2
1.26
44.5
Dolly Varden EO/M
Esox
34.0
NA
21.7
Directly determined EC24
Cyprinodon
27.0
0.94
28.7
Desert pupfish EO/WB (1.20)
/all fishM/WB (1.27)
0. mykiss
24.5
1.92
12.8
Rainbow trout EO/M
Rudolph et al.
2008
24.7
NA
16.6
Directly calculated EC10
Nautilus 2011
27.7
1.81
15.3
Cutthroat trout EO/M
0. clarkii
26.2
NA
16.0
Geometric mean of two studies
Oncorhynchus
25.3
NA
14.3
Geometric mean of two SMCVs
Micropterus
26.3
1.19
22.2
Micropterus EO/M
Salmo
21.0
1.14
18.5
Brown trout EO/WB (1.45)
/all fish M/WB (1.27)
52

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Taxon
i:o
Chronic
Value
KO/.M
(1
Direct or
Calculated
Muscle Uepro
Chronic Value
Direct Calculation or
Basis lor KO/.M CI
(from Appendix 1$)
Coyle et al.
1993
26.3
1.38
19.1
Bluegill sunfish EO/M
Doroshov et al.
1992a
22.6
NA
15.7
Directly calculated ECio
Hermanutz et
al. 1992, 1996
14.7
NA
13.4
Directly calculated ECio
Lepomis
20.6
NA
15.9
Geometric mean of three studies
Acipenser
15.6
NA
11.9
Directly calculated ECio
Table 3.7. The Lowest Four Rcproductivc-F-nect Fish Aluscle GAICVs.
Relative Sensitivity
Knnk
(ienus
GM( V
(nig Se/kg dw muscle)
4
Salmo
18.5
3
Lepomis
15.9
2
Oncorhynchus
14.3
1
Acipenser
11.9
Because the factors used to convert egg-ovary to muscle concentrations vary across
species based on empirical data, the whole-body rankings differ from both from the egg-ovary
rankings and the muscle rankings. With N=15 GMCVs, the 5th percentile projection yields a
muscle criterion element concentration of 11.3 mg Se/kg dw muscle, lower than the muscle value
for the most sensitive fish genus tested, Acipenser. Figure 3.3 compares the fish muscle criterion
element concentration to the distribution of fish muscle reproductive chronic values in Table 3.6.
53

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- X
+
I
10
Rank
A
Salvelinus
X
Cyprinodon
+
Micropterus

Esox
X
Gambusia
-
Salmo
~
Lepomis
~
Oncorhynchus

Acipenser
	
M FCV
Figure 3.3. Distribution of Reproductive-Effect GMCVs for Fish, either Measured as
Muscle in the Original tests, or Measured as Egg-Ovary Concentrations but Converted to
Muscle Concentrations.
(As shown in Table 3.6). (Live-bearer Gambusia was converted from WB to muscle).
3.1.4 Invertebrate Chronic Effects
Below is a brief synopsis of the experimental design of the available invertebrate chronic
toxicity tests, and the resulting chronic values.
Brachiomts calyciflorus (rotifer)
Dobbs et al. (1996) exposed Brachiomts calyciflorus to selenate in natural creek water for
25 days in a three-trophic level food chain test system. This is one of two laboratory-based
experiments (also see Bennett et al. 1986) that involved exposing algae to selenium (in this case
as sodium selenate) in water, and subsequently feeding the algae to rotifers which were in turn
fed to fish (fathead minnows). In the Dobbs et al. (1996) study, the rotifers and fish were
exposed to the same concentrations of sodium selenate in the water as the algae, but consumed
selenium bioaccumulated in the next lower trophic level. Rotifers did not grow well at
concentrations exceeding 108.1 |ig Se/L in water, and the population survived only 6 days at
selenium concentrations equal to or greater than 202.4 |ig Se/L in the water (40 |ig Se/g dw in
the algae). Regression analysis of untransformed growth data (dry weight), determined 4 day
post-test initiation, resulted in a calculated ECio of 37.84 |ig Se/g dw tissue.
54

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Lumbriculus variegatus (oligochaete, blackworm)
Although not intended to be a definitive toxicity study for blackworms, Besser et al.
(2006) evaluated the bioaccumulation and toxicity of selenized yeast to the oligochaete,
Lumbriculus variegatus, which was intended to be used for dietary exposure in subsequent
studies with the endangered desert pupfish, Cyprinidon macularius. Oligochaetes fed selenized-
yeast diets diluted with nutritional yeast (54 to 210 mg Se/kg) had stable or increasing biomass
and accumulated Se concentrations as high as 140 mg/kg dw. The oligochaetes fed the undiluted
selenized-yeast (826 jag/g Se dry wt.) showed reduced biomass. The effect level is considered
>140 mg Se/kg dw.
Centroptilum triangulifer (mayfly)
Mayfly larvae (Centroptilum triangulifer) were exposed to dietary selenium contained in
natural periphyton biofilms to eclosion (emergence) (Conley et al. 2009; Conley et al. 2011;
Conley et al. 2013). In Conley et al. (2009), the periphyton fed to the mayfly larvae were
75
exposed to dissolved selenite (radiolabeled Se) in November 2008 (12.6 and 13.9 |ig/L) and in
January 2009 ( 2.4, 2.4, 4.9, 10.3, and 10.7 |ig/L). Periphyton bioconcentrated selenium an
average of 1113-fold over the different aqueous selenium concentrations (see Table E-22 in
Appendix E). Twenty 4 to 6-day old mayfly larvae were exposed for 4.5 to 6 weeks to each of
the periphyton diets until the larvae eclosed to subimagos (final pre-adult winged stage). The
subimagos were allowed to emerge to the adult imago stage which deposited their egg masses in
Petri dishes. Selenium concentrations were measured in postpartum adults along with their dry
weights and clutch size. Selenium increased in concentration from periphyton to the adult
mayflies (trophic transfer factor) an average of 2.2-fold. The authors observed a reduction in
fecundity with diets containing more than 11 mg Se/kg dw, which is considered the dietary
threshold for this study. Using the trophic transfer factor of 2.2, the periphyton selenium
concentration of 11 mg/kg dw translates to an adult mayfly selenium concentration of 24.2
mg/kg dw.
Conley et al. (2011) exposed larval C. triangulifer larvae similar to Conley et al. (2009)
to two different rations of periphyton (lx and 2x) containing low, medium and high selenium
levels to evaluate the effect of feeding ration on the bioaccumulation of selenium and life cycle
performance of the mayfly. Mayfly larvae were fed either a lx or 2x ration of periphyton loaded
with the three different selenium levels until the larvae eclosed to subimagos after 25-29 days.
55

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Average periphyton Se concentrations for the three treatments in the lx ration study were 4.2,
11.9, and 27.2 mg/kg dw, respectively. In the 2x ration study, average periphyton concentrations
for the three treatments were 9.5, 19.9, and 40.9 mg/kg dw, respectively (Conley et al. 2011).
Subimagos were induced to emerge to adults in petri dishes and their clutch size measured
through digital imaging. Mayflies fed the lx ration had 54% and 72% reductions in survival
relative to controls in the medium and high Se treatment levels, respectively, both of which were
significant (p<0.05). The mayflies fed the lx ration also had significant reductions in fecundity
in the low (44% reduction), medium (63% reduction) and high (77% reduction) Se treatment
levels. However, for the mayflies that were fed the 2x ration, there were no significant
differences between the controls and any of the three Se treatment levels for any of the endpoints
measured including survival and fecundity. The 2x ration mayflies had 60% more biomass than
the lx ration mayflies. This growth difference explains why the lx ration mayflies had higher
concentrations of Se in their tissues (see Table E-23 in Appendix E). The two different rations
resulted in vastly different effect levels for Se, <12.8 mg/kg dw in the lx ration test and >37.3
mg/kg dw in the 2x ration. It is apparent from this study that if the mayflies do not obtain
sufficient nutrition, they are more sensitive to selenium. Although reduced feeding levels occur
in nature, it is a confounding variable in this study that cannot be used to set a chronic effect
level for selenium.
Conley et al. (2013) evaluated the accumulation of selenite and selenate into periphyton
with a subsequent feeding exposure to mayfly larvae. As in the previous studies, C. triangulifer
larvae were fed periphyton previously exposed to different concentrations of selenium. In this
study, periphyton plates were first exposed to low (10 |ig/L) and high (30 |ig/L) concentrations
of either selenite or selenate and then fed to mayfly larvae to eclosion and to subimagos. The
mean concentrations of selenium in the periphyton fed to the mayflies were 2.2, 12.8 and 37
mg/kg Se dw in the control, low and high treatments, respectively. Mayfly tissue (subimago)
concentrations (extrapolated from Figure 4a in Conley et al. 2013) were approximately 4-7, 20-
35, and 45-75 mg/kg Se dw, in the control, low and high treatments, respectively. The authors
reported significant reductions in survival from the control in the high Se treatment (both pooled
data and individual selenite and selenate treatments), but no significant differences were
observed in the low Se treatments. Secondary production (mayfly biomass) was significantly
reduced relative to the control in the high Se treatment for both selenium species. For the low Se
56

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exposure treatments, secondary production was not significantly different than the control for the
selenite treated periphyton exposure, but was for the selenate and pooled data suggesting an
effect level between 20 and 35 mg/kg Se dw. These results as well as those observed in 2x ration
exposures in Conley et al. (2011) where no effects were observed at 37.3 mg/kg Se dw generally
support the chronic value determined for Conley et al. (2009) of 24.2 mg/kg Se dw. This
information included tabulated data from these studies presented in Appendix E.
3.1.5 Summary of Relevant Invertebrate Tests
The available measured invertebrate whole-body effect concentrations are shown in
Table 3.8. Because the intent of this assessment is to derive a concentration expressed in terms
of fish tissue, Table 3.8 also provides information on how concentrations in invertebrate tissue
are translated (in Section 3.2) across media to predicted WB fish tissue concentrations (Trophic
Level 3, TL3) in a system having invertebrates and fish. That is, consistent with the
bioaccumulation modeling approach of Section 3.2, the second column of Table 3.8 uses the
median trophic transfer factor of 1.21 from Table 3.11 to yield expected WB fish tissue
concentrations in a system having invertebrates and fish. Whether comparing TL2 (invertebrate)
whole-body GMCVs directly to Table 3.4 TL3 (fish) whole-body GMCVs, or via the trophic
transfer adjustment in the second column of Table 3.8, it is apparent that invertebrates are not
among the most sensitive species.
The relative insensitivity of invertebrates overall when compared with the fish whole-
body concentrations demonstrates that invertebrates are expected to be generally protected by
selenium criterion values derived from fish. It should be noted that mayflies appear to be the
most sensitive invertebrate group tested; their whole-body effects levels just below the least
sensitive fish genus (Salvelinus, Dolly Varden) on whole-body basis. However these mayfly
results have some uncertainty due to the indicated effect of feeding ration on selenium toxicity to
mayflies that has not been fully defined. The rotifer and lumbricuius tests indicate that these
organisms are less sensitive than any tested fish genus on a whole-body basis. Therefore, the
invertebrates are considered implicitly in the species sensitivity distribution, and are counted
toward the number of values available to calculate the fish tissue criterion elements (as egg-
ovary, whole-body, and muscle), and the missing invertebrate MDRs (4 and 5) are considered
satisfied by the available invertebrate data.
57

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Table 3.8. Ranked Invertebrate Whole-Body Chronic Values with Translation to Expected
Accompanying Fish Whole-Body Concentrations		
SMCV & (i.MCY
as measured
( Trophic Level 2)
(mg Se/kg d\v WIS)
Accompanying Trophic Level 3
Median Whole-Body Concentration
Predicted by liioacciimiilalion
Model (Section 3.2)
(mg Se/kg d\v WIS I I.3)
Species
> 140
> 169.4
Oligochaete, black,
Lumbriculus variegatus
37.84
45.8
Rotifer,
Brachionus calyciflorus
24.2
29.3
Mayfly,
Centroptilum triangulifer
3.1.6 Selenium Fish Tissue Toxicity Data Fulfilling Minimum Data Needs
The toxicity data currently available for genera and species fulfilling the EPA Ambient
Water Quality Criteria Guidelines recommendations for calculation of the freshwater chronic
criterion are described in Sections 3.1.1, 3.1.4, 6.1.2 and Appendix C, and are summarized in
Table 3.9
Table 3.9. Minimum Data Requirements Summary Table Reflecting the Number of Species
and Genus Level Mean Values Represented in the Chronic Toxicity Dataset for Selenium
in Freshwater.

(ienus Mean Chronic
Species Mean Chronic
Kreshwater .Minimum Data Requirement
Value (G.MCV)
Value (SMCV)
1. Family Salmonidae in the class Osteichthyes
3
4
2. Second family in the class Osteichthyes,
preferably a commercially or recreationally
2
2
important warmwater species


3. Third family in the phylum Chordata (may be
in the class Osteichthyes or may be an
amphibian, etc.)
5
5
4. Planktonic Crustacean
See text
See text
5. Benthic Crustacean
See text
See text
6. Insect
1
1
7. Family in a phylum other than Arthropoda or
Chordata (e.g., Rotifera, Annelida, or
1
1
Mollusca)


8. Family in any order of insect or any phylum
not already represented
1
1
Total
15
16
58

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The first three of these MDRs in Table 3.9 are easily fulfilled by the fish species
represented in Sections 3.1.1, 6.1.2 and Appendix C. Because the field observations of
contaminated sites have found effects on fish and birds in the absence of changes in invertebrate
assemblages, scientific studies on the chronic toxicity of dietary selenium for invertebrates has
been very limited. The few dietary chronic toxicity studies that are available for invertebrate
species (arthropods , rotifers, and worms) indicate that they are generally less sensitive than fish,
with available data indicating invertebrate whole body mean chronic values ranging from
approximately 3 to 12 times higher than the fish whole body criterion element value
recommended in this document (Section 3.1.4), The above invertebrate data address MDRs 6-8,
leaving only MDRs 4 and 5, for the planktonic and benthic crustaceans, to be addressed. Because
the 5th percentile calculation methods for the FCV use actual numeric values for the GMCVs of
the four most sensitive (fish) genera in the selenium dataset, it is only necessary to know that the
more tolerant genera have GMCVs that are greater than those of the lowest four. A
recommendation in the draft white paper on Aquatic Life Criteria for Contaminants of Emerging
Concern Part I (U.S. EPA 2008b), which was supported by the Science Advisory Board, states
"because only the four most sensitive genus mean chronic values (GMCVs) are used in the
criterion calculations, chronic testing requirements for a taxon needed to meet an MDR should
be waived if there is sufficient information to conclude that this taxon is more tolerant than the
four most sensitive genera."
Currently, there are no available data on the chronic toxicity to crustaceans via dietary
exposure to selenium. Since there are data available for insects (Centroptilum spp. mayfly), EPA
used the taxonomic association at the level of phylum (Arthropoda) to allow insects to be a
surrogate for crustaceans. There is also associative evidence that macroinvertebrates in general
are less sensitive than fish. At sites where there have been documented effects to fish and
aquatic-dependent birds from selenium exposure (e.g., Kesterson Reservoir, Belews Lake, Hyco
Reservoir), field observations and data indicate that there has been no evidence of effects to
macroinvertebrates including crustaceans (Janz et al. 2010). In addition, Janz et al. (2010) notes
that the key vector for selenium toxicity via maternal transfer is selenium loading in the egg via
vitellogenesis. Crustaceans, and other arthropods are not known to deposit a significant amount
of vitellogenin in the egg compared with oviparous vertebrates like fish, therefore, less selenium
is likely transferred to the egg via deposition of vitellogenin. These mechanistic considerations
59

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are thus consistent with the absence of observed field effects on aquatic macroinvertebrates,
including crustaceans and other arthropods, and with the Chapman et al. (2009, 2010) expert
consensus that it is the egg-laying vertebrates that are most at risk.
Applying this concept to the selenium criterion 5th percentile calculations, GMCVs for
MDRs 4 and 5 (the two crustacean MDRs) should be waived and counted in the total number of
GMCVs in the dataset, based on (a) the difference in the measured effect values discussed above,
and (b) the lack of observed invertebrate field effects linked to selenium (for example, as
concluded by Lemly 2002, pages 21-23, and Janz et al. 2010). Thus data are adequate to fulfill
the data needs for developing a chronic selenium criterion.
The total number of GMCVs available to derive the chronic criterion is 15. These include
ten fish genera from Sections 3.1.1 and 6.1.2 (Acipenser, Salmo, Lepomis, Micropterus,
Oncorhynchus, Pimephales, Gambusia, Esox, Cyprinodon, and Salvelinus) [Added to these are
the tested invertebrate genera Centroptilum, Brachionus, and Lumbriculus from Section 3.1.4,
and lastly the two waived genera for MDRs 4 and 5 (crustaceans).
3.2 Chronic Water Column-based Selenium Criterion Element
3.2.1 Translation from Fish Tissue Concentration to Water Column Concentration
EPA derived the chronic water column selenium criterion element by translating the egg-
ovary concentration to an equivalent water concentration. EPA worked with USGS to derive a
translation equation that utilizes a mechanistic model of bioaccumulation previously published in
peer-reviewed scientific literature (Luoma et. al. 1992; Wang et. al. 1996; Luoma and Fisher
1997; Wang 2001; Schlekat et al. 2002b; Luoma and Rainbow 2005; Presser and Luoma 2006,
2010; Presser 2013). This model quantifies bioaccumulation in animal tissues by assuming that
net bioaccumulation is a balance between assimilation efficiency from diet, ingestion rate, rate of
direct uptake in dissolved forms, loss rate, and growth rate. The basic model is given as:
60

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c,
tissue
\(K'C^)+(AEyIRyClx,,l
(k,+g)
(Equation 1)
Where:
IR
water
tissue
g
AE
Concentration of chemical in water (|ig/L)
Average concentration of chemical in all tissues at steady-state (|ig/g)
Efflux rate (/d)
Growth rate (/d)
Uptake rate (L/g-d)
Assimilation efficiency (%)
Ingestion rate (g/g-d)
Concentration in food (|ig/g)
3.2.1.1 Simplifying the Bioaccumulation Model
Specific application to selenium bioaccumulation permits the simplification of Equation 1
in two ways. One simplification is removing the parameter representing growth rate (g), and the
other simplification is removing the parameter representing direct aqueous uptake (ku).
Growth Rate
The growth rate constant g is included in Equation 1 because the addition of body tissue
has the potential to dilute the concentration of bioaccumulative chemicals when expressed as
chemical mass per tissue mass. For very hydrophobic chemicals with low excretion rates such as
polychlorinated biphenyls, growth can be an important factor in bioaccumulation estimates
(Connolly and Pedersen 1988). However, Luoma and Rainbow (2005) suggest that for selenium,
growth rate is a relatively inconsequential parameter under most circumstances. Food
consumption is typically high during periods of high growth rate. Because food consumption is
the primary route of selenium uptake in aquatic organisms (Ohlendorf et al. 1986a, b; Saiki and
Lowe 1987; Presser and Ohlendorf 1987; Lemly 1985a; Luoma et al. 1992; Presser et al. 1994;
Chapman et al. 2010), high consumption rates of selenium-contaminated food may counteract
the selenium dilution that occurs with the addition of body tissue during periods of fast growth.
EPA evaluated the effect of removing the parameter g in the Equation 1 by performing a
sensitivity analysis. EPA analyzed a series of hypothetical tissue concentration estimates using
Equation 1 with g ranging between 0 (no growth) and 0.2/day (a relatively high rate of growth).
61

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In one analysis, tissue concentrations of selenium were estimated using static values of IR. In a
second analysis, tissue concentrations of selenium were estimated using values of IR that were
adjusted for growth rate using a method similar to the approach used in a model of organic
chemical accumulation in aquatic food webs (Thomann et al. 1992). As expected, estimates of
selenium tissue concentrations were significantly reduced at progressively higher growth rates
when IR remained constant. However, selenium concentrations remained fairly steady or slightly
increased with progressively higher growth rates when IR was adjusted for the bioenergetics of
growth. This analysis supports the hypothesis that a higher IR (and consequently greater rate of
selenium ingestion) associated with the higher bioenergetic requirements of rapidly growing
young fish tends to oppose the dilution of selenium in their tissues due to growth, whereas a
lower IR (and consequently lower rate of selenium ingestion) associated with the lower
bioenergetic requirements of slower growing older fish tends to oppose the bioconcentration of
selenium in their tissues. EPA concludes from this analysis that omitting the growth rate
parameter g is an appropriate simplification of Equation 1. A more detailed description of this
sensitivity analysis is provided in Appendix J.
Uptake Rate
The uptake rate constant ku is included in Equation 1 to account for direct absorption of
bioaccumulative chemicals in the dissolved phase. However, dietary intake of selenium is the
dominant source of exposure, suggesting that ku may also be relatively inconsequential for
selenium accumulation (Luoma and Rainbow 2005). Because aqueous uptake of selenium makes
up a small percentage of bioaccumulated selenium (Fowler and Benayoun 1976; Luoma et. al.
1992; Roditi and Fisher 1999; Wang and Fisher 1999; Wang 2002; Schlekat et. al. 2004; Lee et.
al. 2006), Presser and Luoma (2010a, b, 2013) deemed removal of ku from Equation 1 as an
acceptable simplification.
EPA evaluated the effect of removing the parameter ku in the Equation 1 by performing a
sensitivity analysis. EPA analyzed a series of tissue concentration estimates using Equation 1 and
a realistic range of ku values for trophic level 2 and trophic level 3 organisms. The analysis
suggests that approximately 75% of selenium exposure in trophic level 2 organisms
(invertebrates) and over 90% of selenium exposure in trophic level 3 organisms occurs through
consumption of selenium-contaminated food. EPA concluded that omitting the aqueous uptake
62

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rate constant ku is an appropriate simplification of Equation 1. A more detailed description of this
sensitivity analysis is provided in Appendix J.
Derivation of the Translation Equation
Disregarding growth (g) and uptake of selenium dissolved in water (ku x Cwater), Equation
1 becomes Equation 2 (Reinfelder et al. 1998):
c _ AExIRxCfood
or:
_ AExIR
tissue	i	food
e	(Equation 2)
Because application of the bioaccumulation model applies to a single species, the
AExIR
combination of species-specific physiological parameters expressed as ;	 remains
ke
AExIR
constant for the species. Thus the EPA defines the expression ;	 as a single species-
ke
specific Trophic Transfer Factor (TTF) given as Equation 3 (Reinfelder et al. 1998):
AExIR
TTF
^e	(Equation 3)
AExIR
Substituting TTF for :	 in Equation 2 yields:
Ctusue = TTF x Cfood	(Equation 4)
The trophic level of the organisms considered can be denoted by superscripts given as:
s~iTL2 	 TTJi ^^ x
ussm	food	(Equation 5)
63

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CTt^ue as defined here represents the steady-state proportional concentration of selenium in the
tissue of trophic level 2 organisms relative to the concentration of selenium in their food source.
Using the same rationale, the average concentration of selenium in the tissues of trophic
level 3 organisms can be expressed as the concentration of selenium in its food multiplied by a
TTF which is given as:
rTL3 = ttftl3 x rTL3
name	food	(Equation 6)
TZ/3	TL 2
For trophic level 3 organisms that consume trophic level 2 organisms, CJood = Ctissue.
Thus:
cIifsue= TTFTL3 X Cjtfsue	(Equation 7)
Substituting c"ue in Equation 7 with TIT'112 x C food in Equation 5 yields:
^TL3 	 rjirrij^TL3 rjirjij^TLl S~lTL2
tissue ~	Jood	(Equation 8)
Defining the term Cue as the concentration of selenium in fish tissue, defining the term
ctifsue as the concentration of selenium in living and nonliving particulate material ingested by
invertebrates, and expressing the product of all TTF values as a single term results in the
equation:
^whole-body = TTFcomposite X Cparticulate	(Equation 9)
where:
Cparticulate = the concentration of selenium in particulate material
Cwhoie-body = the concentration of selenium in the whole body of fish
TTpcompo^te =	product of all trophic transfer factor values
Equation 9 quantitatively expresses selenium bioaccumulation in fish (Cwhoie-body) as the
product of the concentration of selenium at the base of the food web (CparfiCH/ate) and a parameter
representing the trophic transfer of selenium through all dietary pathways	This
model of bioaccumulation is conceptually similar to the model of bioaccumulation utilizing a
bioaccumulation factor (BAF). A BAF is the ratio of the concentration of a chemical in the tissue
64

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of an aquatic organism to the concentration of the chemical dissolved in ambient water at the site
of sampling (U.S. EPA 2001c). Similar to the term /"///x""/x',v"':', a BAF quantitatively represents
the relationship between the chemical concentrations in multiple environmental compartments.
However, a BAF is empirically derived from site-specific measurements, whereas jjpomPslte js
derived from knowledge of the ecological system. Because each TTF is associated with a
particular taxon, jjpomPslte can be inferred for an aquatic system using existing knowledge and
reasonable assumptions, without the considerable time and cost of collecting and analyzing
tissue and water samples.
Equation 9 characterizes the bioaccumulation of selenium as a combination of TTF
parameters from all steps in the dietary pathway of the predator species of interest. Thus it is
possible to differentiate bioaccumulative potential for different predator species and food webs
by modeling different exposure scenarios. For example, where the fish species of interest is a
trophic level 4 predator that primarily consumes trophic level 3 fish, the term	can be
represented as the product of all TTF parameters that includes the additional trophic level given
as:
TTFcompoS1te = TTFTL4 ^	^ y.yj/TLZ	(EqUatiOIl 10)
where:
J"II''112	=	the trophic transfer factor of trophic level 2 species
l'1'l'"3	=	the trophic transfer factor of the trophic level 3 species
TTFTL4	=	trophic transfer factor of the trophic level 4 species
j'!],composite	=	procjuct 0f all trophic transfer factors
Similarly, the consumption of more than one species of organism at the same trophic
level can also be modeled by expressing the TTF at a particular trophic level as the weighted
average of the T'/T's of all species consumed given as:
TTFTLx =  (TTFtTLx X w,)	(Equation 11)
where:
TTFi = the trophic transfer factor of the ith species at a particular trophic level
w;	= the proportion of the ith species consumed
65

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These concepts can be used to formulate an expression of	to model selenium
bioaccumulation in ecosystems with different consumer species and food webs. Figure 3.4
describes four example food web scenarios and the formulation of	to model selenium
bioaccumulation in each of them.
The parameter jjpomPslte quantitatively represents all dietary pathways of selenium
exposure for a particular fish species within an aquatic system. The parameter is derived from
species-specific TTF values representing the food web characteristics of the aquatic system, wt,
the proportion of species consumed. See text for further explanation.
66

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A)	Three trophic levels (simple):
rprpjjcomposite 	 rjirji-piTL3 ^ rjirjijjTLl
TTF713	T7I'TL2
<	 <> <	 SR
B)	Four trophic levels (simple):
y 7'Pcomposite 	4 ^	^ yy 'j_^'TL2
yy/,77.7	yyyTLJ	yy/,77.2
~	4> ~
C> Three trophic levels (mix within trophic levels):
jjpcon.posite = TTFTL3 x [(77^312 x WJ+	x WJ]
Til''1'13 |	 W7	/ r7F'
^2	^	^ ' *
TTF'_
D)	Three trophic levels (mix across trophic levels):
T]Fcon,polite = (pjpTLl x WJ+	x	x
77V ,	li	
:^iv
<	
TTF
E)Four	trophic levels (mix across trophic levels):
jjpcomposite = ^JpTM ^ jjpTLl ^	^jyM x wJ]x jjpTLl
J"lpTL4	]-]],-TI.S	TTJJTL2 /Jtk+
V
W1	^	^ ^ VNf*
I
V Wi
TTFVL4
Figure 3.4. Example Aquatic System Scenarios and the Derivation of the Equation
Parameter TTF0"^8.
Because EPA's objective is to derive an equation that translates a fish tissue
concentration of selenium to a water column concentration, the term Cwater is reintroduced into
67

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Equation 9 by defining the enrichment function EF representing the steady state proportional
bioconcentration of dissolved selenium at the base of the aquatic food web given as:
EF = cvarticulate	(Equation 12)
water
Where:
Cparticuiate = Selenium concentration in particulate material (|ig/g)
CWater = Concentration of selenium dissolved in water (|ig/L)
EF	= Enrichment function (L/g)
Rearranging the terms of Equation 12:
Cparticulate ~ EF x Cwater	(Equation 13)
Substituting EF X Cwater for Cparticuiate in Equation 9 results in:
CM.body = TTFcom^ x EF x CWflter	(Equation 14)
Solving for the concentration of selenium in water in Equation 14 results in:
C
whole-body
 = jjpcomroai, y F	(Equation 15)
Because Equation 15 relates a concentration of selenium in water to the concentration of
selenium throughout all tissues of the body, Cwhoie-body must be converted to an equivalent
concentration in eggs or ovaries. EPA achieved this conversion by incorporating a species-
specific conversion factor (CF) into Equation 15. CI'' represents the species-specific proportion
of selenium in egg or ovary tissue relative to the concentration of selenium in all body tissues
and is given as:
C
QP _ egg-ovary
r
whoh-body	(Equation 16)
Where:
CF	= Whole-body to egg-ovary conversion factor (dimensionless ratio).
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Cegg-ovary = Selenium concentration in the eggs or ovaries of fish (|ig/g)
Cwhole-body = Selenium concentration in the whole body of fish (|ig/g).
Rearranging the terms of Equation 16 yields:
C
egg-ovar y
CF	(Equation 17)
C
ggg	o vEr y
Substituting Cwhoie-body in Equation 15 with	 yields the translation equation:
CF
r
whole-body
r
^ _	egg-ovar y
^ Mint
Yjp composite ^ ^ Qp
(Equation 18)
where TTF comPslte equals the product of all trophic transfer factors from trophic level 2 through
the target fish species.
Equation 18 describes an ecosystem-dependent relationship between the concentration of
selenium in the eggs and ovaries of fish with the concentration of selenium in the water column.
This approach explicitly recognizes the sequential transfer of selenium between environmental
compartments (water, particulate material, invertebrate tissue, fish tissue, and eggs and/or ovary
tissue) by incorporating quantitative expressions of selenium transfer from one compartment to
the other. Because this approach uses food web modeling along with species-specific TTF and
CF parameters to quantify most of the transfer between compartments, the only field
measurements needed to relate selenium in egg-ovary and water are measurements from the
water column and particulate material sufficient to calculate EF.
3.2.2 Equation Parameters
Empirical or laboratory data related to selenium bioaccumulation in aquatic organisms
are needed to derive the equation parameters EF, TTF, and CF. EPA obtained data from
published literature as described above The search resulted in the retrieval of 63 acceptable
studies containing a total of 8,707 selenium measurements at 768 aquatic sites (2,927 from
water, 373 from algae, 29 from detritus, 821 from sediment, 1,324 from various species of
69

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invertebrates, and 3,233 from various species of fish) and 34 acceptable studies yielding 139
physiological constants (48 values of ke, 81 values of AE, and 10 values of IR). EPA used this
collection of selenium measurements to calculate site-specific EF values and develop species-
specific TTF and CF values in an unbiased and systematic manner. A more detailed description
of how EPA calculated EF is described below. How EPA calculated TTF and CF is described in
detail in Appendix B.
3.2.2.1 Derivation of Trophic Transfer Factor (TTF) Values
EPA derived TTF values for taxonomic groups of invertebrates and fish by either using
physiological coefficients found in the literature, or by evaluating the empirical relationship
between matched pairs of selenium measurements in organisms and the food they consumed.
When physiological coefficients were available, EPA calculated a TTF value using Equation 3
(Section 3.2.1):
TTF=^
^e	(Equation 3)
Where:
^e	= Elimination rate constant (/d)
AE = Assimilation efficiency (%)
IR = Ingestion rate (g/g-d)
EPA also derived TTF values using empirical measurements of selenium from field
studies. EPA searched its collection of available selenium measurements and identified
measurements taken from aquatic organisms. For each measurement from an aquatic organism,
EPA searched for additional measurements from other aquatic organisms or particulate material
that was collected from the same aquatic site and of a type deemed likely to be ingested as a food
source or in conjunction with feeding activity (i.e., a lower trophic level). If multiple lower
trophic level measurements were matched to an aquatic organism measurement, the median of
the lower trophic level measurements was calculated. Each pair of measurements, one taken
from an aquatic organism and the other taken from lower trophic level organisms or particulate
material, was designated as a matched pair. EPA limited particulate data used to calculate
70

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invertebrate TTFs from field data to those aquatic sites with at least two particulate selenium
measurements paired with invertebrate selenium measurements, and only used sediment
measurements if there was at least one measurement from algae or detritus. If selenium
concentrations from more than category of particulate material (algae, detritus, or sediment) were
available, EPA used the median Se concentration of the available categories as the particulate
concentration for that site.
Because selenium is transferred to aquatic animals primarily through aquatic food webs,
the observable concentration of selenium in different environmental compartments may vary
over time. To establish an appropriate time period with which to define matched pairs of
selenium measurements, the effect of sample collection time on the relationship between
selenium concentrations in different media was analyzed. EPA defined matched pairs of
selenium measurements as described above using different relative collection time ranges and
estimated the strength of the relationship between the two measurements by calculating the
Pearson product-moment correlation coefficient (r). Figure 3.5 shows the correlation coefficients
for selenium measurements taken from the same aquatic sites when the measurement collection
times were systematically shifted relative to one another. Each correlation coefficient was
calculated from a set of data within a specified range of relative collection times with respect to
the higher trophic level. For example, the correlation coefficient calculated from invertebrate and
fish measurements with a relative sample collection time of 30 to 60 days were from invertebrate
and fish samples collected at the same site, with the fish samples collected 30 to 60 days after the
invertebrate samples. Similarly, the correlation coefficient calculated from invertebrate and fish
measurements with a relative collection time of -60 to -30 days were from invertebrate and fish
samples that were collected at the same site, with the fish samples collected 30 to 60 days before
the invertebrate samples.
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Particulate versus invertebrate
1
#   
+ i
i
i
i


{ "
1 \
i
i
i
i J
~

-1
Invertebrate versus fish
i i    I
I

i  ~
.
-1
Uj Uj
I-Ix-e-s-
2 if
c	_ o c c s o o
ft <5:
o o
a s
I
o  o 8 o 8
8 8
oooooooooooooooooooo
DOOOOOOOOOOOOO
8 s s a s 
o o o o o
Number of days between collection time of paired samples
Figure 3.5. Effect of Relative Sample Collection Time on Correlation Coefficients of
Selenium Measurements in Particulate Material, and Invertebrate and Fish Tissue.
Error bars indicate the 95% confidence interval of r calculated using Fisher's r to z
transformation. Horizontal dashed line indicates r = 0; vertical dashed line indicates relative
collection time expected to have the highest correlation. The absence of a correlation coefficient
indicates an insufficient quantity of data at the specified relative collection time range.
The results of this analysis suggest that the relationship between selenium concentrations
in particulate material and invertebrate tissue and between invertebrate tissue and fish tissue is
insensitive to relative collection time within a one year time period. These results also suggest
that selenium becomes relatively persistent in the aquatic ecosystem once dissolved selenium
transforms to particulate selenium and becomes bioavailable. On the basis of these analyses,
EPA concludes that selenium measurements from samples collected at the same aquatic site
within one year of each other are acceptable to use as matched pairs of measurements from the
aquatic sites. Note that EPA chose a relative collection time period of one year on the basis of
data taken from many different aquatic sites. Individual aquatic sites may have selenium loads
and/or bioaccumulation characteristics that require different relative collection time criteria to
accurately characterize selenium relationships.
After matched pairs of selenium measurements from samples collected in the field were
identified, EPA evaluated two different analytical approaches to derive species-specific TTF
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values from them. TTF was previously defined above as the steady state proportion relating the
concentration of selenium in the tissue of aquatic organisms to the concentration of selenium in
the food they ingest such that:
Ct^ue = TTF x Cfood	(Equation 4)
Rearranging the terms of Equation 4 yields Equation 19:
C
yyy	 tissue
~ r
fod	(Equation 19)
Because TTF can be defined as the ratio of the concentration of selenium observed in the
tissue of an aquatic organism to the concentration of selenium observed in the tissue or material
the organism ingests, one approach for deriving TTF values from field data is to simply use the
ratio of the two values. EPA evaluated this approach by calculating the ratios for all matched
pairs of selenium measurements, and for each species or taxonomic group, used a statistic of
central tendency of the distribution of ratios as the TTF value. An advantage of quantifying the
relationship between selenium in two environmental compartments using ratios is that it is a
simple and straightforward method that is conceptually similar to a bioaccumulation factor
(BAF). A disadvantage of this approach is that it presumes that the quality and quantity of data
used to derive the ratios adequately represent the relationship being characterized. Furthermore,
many aquatic organisms tend to bioaccumulate more metals at low environmental concentrations
(McGeer et al. 2003; Borgman et al. 2004; DeForest et al. 2007; U.S. EPA 2007). Thus a
distribution of ratios could be biased toward larger values if the data are obtained from aquatic
systems with low selenium concentrations.
Another analytical approach for deriving TTF values from matched pairs of selenium
measurements is to model the species-specific relationships using linear regression. One
possibility is to regress the concentration of selenium in the food of a particular species or
taxonomic group with the concentration of selenium in the organism's tissue, and use the
regression coefficient as the TTF. EPA evaluated this approach by applying ordinary least
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squares (OLS) linear regression on the matched pairs of data. The regression coefficient (slope of
the fitted line) was then taken as the TTF value for that species or taxonomic group. An
advantage of this regression approach is that it estimates the quantitative relationship of selenium
across a range of environmental concentrations in a manner that allows statistical assessment.
Disadvantages of this regression approach include the assumption that the underlying data are
normally distributed; one or a few very high values can have a disproportionate influence on the
slope of the fitted line; and the bioaccumulation model does not account for a non-zero y-
intercept. Constraining the y-intercept to zero (also known as regression through the origin or
RTO) eliminates the added complexity of a non-zero y-intercept. However, RTO further
increases the disproportionate influence of one or a few high values on the slope of the fitted
line. Furthermore, RTO does not provide a straightforward way of evaluating goodness of fit
(Gordon 1981).
After evaluating both approaches, EPA decided to use a hybrid approach by designating
the median of the ratio of matched pairs of selenium measurements as the TTF value, but only if
OLS linear regression of those data resulted in a significant (P < 0.05) fit and positive regression
coefficient. Requiring a significant positive OLS linear regression coefficient confirms the
relationship between selenium in organisms and the food they ingest is adequately represented
by the available data. Using the median of the individual ratios provides an estimate of central
tendency for that relationship that is less sensitive to potential bias from measurements taken
from aquatic systems with very high or very low selenium concentrations. Some aquatic
organisms exhibit selenium bioaccumulation inversely related to water concentration (McGeer et
al. 2003; Borgman et al. 2004; DeForest et al. 2007). This inverse relationship is likely due to
saturation uptake kinetics of specific transport mechanisms that regulate metals bioaccumulation
within certain ranges (U.S. EPA 2007). EPA evaluated the effect of very high and very low
selenium concentrations on the calculation of TTF values using the hybrid approach described
above by calculating TTF values excluding selenium measurements above 10, 25, 50, and 100
Hg/g and below 0.1, 0.5, 1.0, and 2.0 |j,/g. EPA found that excluding very high or very low
selenium measurements had minor effects on TTF values. EPA concludes that using the median
ratio effectively attenuates effects of selenium concentration on the calculation of TTF values
using the hybrid approach described above and allows the use of all available data without the
need to introduce additional arbitrary exclusion criteria.
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EPA calculated TTF values for 13 invertebrate species and 32 fish species that live in
freshwater aquatic environments in North America. The data used to derive these TTF values are
provided in Appendix B. The final TTF values are listed in Table 3.10 and Table 3.11. The
presence of physiological coefficients for a taxon in Table 3.10 and Table 3.11 indicates that the
TTF values were calculated using those parameters. The absence of physiological coefficients for
a taxon indicates that EPA derived the TTF value using field data. If a TTF value could be
calculated from both physiological coefficients and field data, EPA used the TTF value
calculated from the substantially larger number of field measurements to minimize statistical
uncertainty.
Table 3.10. EPA-Derived Trophic Transfer Factor (TTF) Values for Freshwater Aquatic
Invertebrates.
Com moil name
Scientific name
\i:
IK
K
Ill
Crustaceans
amphipod
Hyalella azteca
-
-
-
1.22
copepod
copepods
0.520
0.420
0.155
1.41
crayfish
Astacidae
-
-
-
1.46
water flea
Daphnia magna
0.406
0.210
0.116
0.74
Insects
dragonfly
Anisoptera
-
-
-
1.97
damselfly
Coenagrionidae
-
-
-
2.88
mayfly
Centroptilum triangulifer
-
-
-
2.38
midge
Chironimidae
-
-
-
1.90
water boatman
Corixidae
-
-
-
1.48
Mollusks
asian clama
Corbicula fluminea
0.550
0.050
0.006
4.58
zebra mussel
Dreissena polymorpha
0.260
0.400
0.026
4.00
Annelids
blackworm
Lumbriculus variegatus
0.165
0.067
0.009
1.29
Other
zooplankton
zooplankton
-
-
-
1.89
a Not to be confused with Potamocorbula amurensis
75

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Table 3.11. EPA-Derived
rophic Transfer Factor (TTF) Values for Freshwater Fish.
Common name
Scientific name
\i:
IK
K
1 IK
Cypriniformes
blacknose dace
Rhinichthys atratulus
-
-
-
0.71
bluehead sucker
Catostomus discobolus
-
-
-
1.04
longnose sucker
Catostomus catostomus
-
-
-
0.90
white sucker
Catostomus commersonii
-
-
-
1.11
flannelmouth sucker
Catostomus latipinnis
-
-
-
0.98
common carp
Cyprinus carpio
-
-
-
1.20
creek chub
Semotilus atromaculatus
-
-
-
1.06
fathead minnow
Pimephales promelas
-
-
-
1.57
red shiner
Cyprinella lutrensis
-
-
-
1.31
redside shiner
Richardsonius balteatus
-
-
-
1.08
sand shiner
Notropis stramineus
-
-
-
1.56
Cyprinodontiformes
western mosquitofish
Gambusia affinis
-
-
-
1.21
northern plains killifish
Fundulus kansae
-
-
-
1.27
Esociformes
northern pike
Esox lucius
-
-
-
1.78
Gasterosteiformes
brook stickleback
Culaea inconstans
-
-
-
1.79
Perciformes
black crappie
Pomoxis nigromaculatus
-
-
-
2.67
bluegill
Lepomis macrochirus
-
-
-
1.03
green sunfish
Lepomis cyanellus
-
-
-
1.12
largemouth bass
Micropterus salmoides
-
-
-
1.39
smallmouth bass
Micropterus dolomieu
-
-
-
0.86
striped bass
Morone saxatilis
0.375
0.335
0.085
1.48
walleye
Sander vitreus
-
-
-
1.60
yellow perch
Perca flavescens
-
-
-
1.42
Salmoniformes
brook trout
Salvelinus fontinalis
-
-
-
0.88
brown trout
Salmo trutta
-
-
-
1.38
mountain whitefish
Prosopium williamsoni
-
-
-
1.38
cutthroat trout
Oncorhynchus clarkii
-
-
-
1.12
rainbow trout
Oncorhynchus mykiss
-
-
-
1.07
Scorpaeniformes
mottled sculpin
Cottus bairdi
-
-
-
1.38
sculpin
Cottus sp.
-
-
-
1.29
Siluriformes
black bullhead
Ameiurus melas
-
-
-
0.85
channel catfish
Ictalurus punctatus
-
-
-
0.68
76

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For fish species without sufficient data to directly calculate a TTF value, EPA estimated
the TTF value by sequentially considering higher taxonomic classifications until one or more
taxa for which a calculated TTF value was available matched the taxon being considered. If the
lowest matching taxon was common to more than one species with a TTF value available, EPA
used the median TTF from the matching species. For example, although data to directly calculate
TTF for Chrosomus eos (northern redbelly dace) were not available, this species is in the family
Cyprinidae, which also includes Rhinichthys atratulus (blacknose dace), Cyprinus carpio
(common carp), Semotilus atromaculatus (creek chub), Pimephales promelas (fathead minnow),
Cyprinella lutrensis (red shiner), Richardsonius balteatus (redside shiner), and Notropis
stramineus (sand shiner). Because Cyprinidae is the lowest taxonomic classification where
Chrosomus eos matches a species with an available TTF value, the median of the blacknose
dace, common carp, creek chub, fathead minnow, red shiner, redside shiner, and sand shiner TTF
values was used as the TTF value for northern redbelly dace. The data and analyses used to
calculate all TTF values including those estimated by taxonomic classification are provided in
Table B-8 of Appendix B.
3.2.2.2 Derivation of Whole-Body to Egg-Ovary Conversion Factor (CF) Values
The parameter CF (conversion factor) in Equation 18 (Section 3.2.1) represents the
species-specific partitioning of selenium as measured in the whole-body and in egg-ovary tissue.
EPA derived species-specific CF values by applying the same method used to derive species-
specific TTF values using empirical measurements of selenium concentrations in different tissues
of the same fish. To derive whole-body to egg-ovary CF values, EPA defined matched pairs of
selenium measurements from the whole-body and from the eggs or ovaries measured from the
same individual fish or from matched composite samples. Egg-ovary concentration was defined
as a measurement from either the eggs or the ovaries. If multiple measurements from both eggs
and ovaries of the same individual or matched composite sample were available, the average
value was used. For example, both egg and ovary measurements were available for five of the 27
egg-ovary concentrations used to calculate the bluegill egg-ovary to whole body CF (Coyle et al.
1993), and 16 of the 69 measurements used to calculate the cutthroat trout egg-ovary to muscle
CF (Kennedy et al. 2000).
Similar to the procedure used to derive TTF values, EPA first confirmed a statistical
relationship between egg-ovary and whole body selenium for each species using OLS linear
77

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regression of the matched pairs of measurements. If the regression resulted in a significant fit
(P<0.05) with a positive regression coefficient, EPA calculated the ratio of the egg-ovary to
whole body selenium concentration of each matched pair and used the median ratio as the CF
value for the species. A detailed comparison of the advantages and disadvantages of the median
ratio and least squares regression approaches to calculating CFs, along with a comparison of CFs
calculated from median ratios, OLS regression following log transformation, and total least
squares (TLS) regression following log transformation is included in Appendix N.
EPA had sufficient egg-ovary and whole-body selenium measurements to directly derive
egg-ovary to whole body CF values for 13 species of fish. However, matched pairs of selenium
measurements in eggs and/or ovaries and muscle tissue, and matched pairs of selenium
measurements in muscle and whole body were also available. To derive CF values for additional
fish species, EPA used either the additional data or a taxonomic classification approach to
estimate CF.
For those species of fish with neither sufficient data to directly calculate an egg-ovary to
whole body CF, nor data to calculate a conversion factor for egg-ovary to muscle or whole body
to muscle, EPA first estimated CF following the approach described above for the estimation of
TTF values. In this first approach, EPA sequentially considered higher taxonomic classifications
until one or more taxa for which a calculated CF value was available matched the taxon being
considered, and if the lowest matching taxon was common to more than one species with a CF
value available, EPA used the median CF from the matching species. For example, CF data are
not available to directly calculate CF for Lepomis microlophus (redear sunfish); however, genus-
level CFs for Lepomis cyanellus (green sunfish) and Lepomis macrochirus (bluegill) are
available. Thus, EPA used the median CF values of Lepomis cyanellus and Lepomis macrochirus
for redear sunfish.
For fish species without sufficient data to directly calculate an egg-ovary to whole body
CF, but which had sufficient data to calculate a conversion factor for either egg-ovary to muscle
or whole body to muscle, EPA followed a two stage approach based on taxonomic similarity. If a
fish species had a species specific egg-ovary to muscle conversion factor, but no whole body
data with which to calculate an egg to whole body CF, available data for other species would be
used to estimate a muscle to whole body conversion factor for that species based on taxonomic
relatedness. The estimated muscle to whole body factor would be multiplied by the directly
78

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measured egg-ovary to muscle factor to estimate an egg-ovary to whole body CF for that species.
For example, rainbow trout has a species specific egg-ovary to muscle conversion factor of 1.92,
but does not have a species specific egg-ovary to whole body CF. Using the taxonomic approach
described above, the most closely related taxa to rainbow trout with muscle to whole body
conversion factors are in the class Actinopterygii. The median conversion factor for the eight
species within that class is 1.27. The final egg-ovary to whole body CF for rainbow trout is 2.44
(Table 3.12), or 1.92 x 1.27.
EPA derived 13 CF values directly from matched pairs of egg-ovary and whole-body
selenium measurements and an additional seven CF values by multiplying EO/M and M/WB
conversion factors. Variability in the CF values for 19 of the 20 fish species was low (Table
3.12). Excluding mountain whitefish, CFs ranged from 1.20 to 3.11, a 2.6-fold difference. CF
variability within each species was also low for 7 of 13 species for which egg-ovary to whole-
body CFs were calculated. The two species with relatively high standard deviations for their CF
values (brown trout and cutthroat trout) contained potentially anomalous hatchery data that
contributed to the variability (see Table 3.12 footnote). These species specific CF values are
listed below in Table 3.12 and in Table B-5 of Appendix B. All CF values including those
estimated using the taxonomic classification approach are provided in Table B-6 in Appendix B.
Table 3.12. EPA-Derived Egg-Ovary to Whole-Body Conversion Factor (CF) Values
Com moil name
Scientific name
CI
Sid. Dev.'1
Acipenseriformes

white sturgeon
Acipenser transmontanus
1.69

Cypriniformes

bluehead sucker
Catostomus discobolus
1.82
0.19
flannelmouth sucker
Catostomus latipinnis
1.41
0.20
white sucker
Catostomus commersonii
1.38
0.36
desert pupfish
Cyprinodon macularius
1.20
0.10
common carp
Cyprinus carpio
1.92
0.49
roundtail chub
Gila robusta
2.07
0.29
fathead minnow
Pimephales promelas
1.40
0.75
creek chub
Semotilus atromaculatus
1.99
1.00
razorback sucker
Xyrauchen texanus
3.11

Esociformes

northern pike
Esox lucius
2.39

Perciformes

bluegill
Lepomis macrochirus
2.13
0.68
79

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Com moil nninc
Scicnlil'ic inline
CF
Slil. Dev.'1
green sunfish
Lepomis cyanellus
1.45
0.23
smallmouth bass
Micropterus dolomieu
1.42
0.19
Salmoniformes

brook trout
Salvelinus fontinalis
1.38

Dolly Varden
Salvelinus malma
1.61

brown trout
Salmo trutta
1.45
1.81"
rainbow trout
Oncorhynchus mykiss
2.44

cutthroat trout
Oncorhynchus clarkii
1.96
2.03b
mountain whitefish
Prosopium williamsoni
7.39

a Standard deviation for CF values for those species that hac
egg-ovary and whole body
concentrations.
b The brown trout and cutthroat trout standard deviations for CF' values of 1.81 and 2.03 are
considerably higher than the other standard deviations in this table. The brown trout data were
taken from two studies, NewFields (2009) and Osmundson et al. (2007). CF values for three of
the four fish samples from Osmundson et al. were four to six times greater than the median.
Also, the NewFields data consisted of samples collected from natural streams and samples
collected from a fish hatchery. The CF values for the fish hatchery samples were four to seven
times lower than the median value. Although collectively, the data set meets the criteria for
including the brown trout CF, the CF values for Osmundson et al. and NewFields hatchery
samples may be anomalously high and low, respectively. Excluding these potentially anomalous
data reduces the brown trout standard deviation to 0.47. The cutthroat trout CF values are from
two sources (Formation 2012 and Hardy 2005). The reason for the higher variability in the
cutthroat trout CF values is due to the relatively higher CF values in the hatchery fish from the
Formation study. The standard deviation for cutthroat trout drops to 0.62 if the hatchery fish are
excluded. See Appendix B for a presentation of the data for both of these species.
3.2.2.3 Calculation of Site-Specific Enrichment Factor (EF) Values
The most influential step in selenium bioaccumulation occurs at the base of aquatic food
webs (Chapman et al. 2010). The parameter EF characterizes this step by quantifying the
partitioning of selenium between the dissolved and particulate state. EF can vary by at least two
orders of magnitude across aquatic systems (Presser and Luoma 2010). The greatest reduction in
uncertainty when translating a fish tissue concentration of selenium to a water column
concentration using Equation 18 is achieved when spatially and temporally coincident site-
specific empirical observations of dissolved and particulate selenium of sufficient quality and
quantity are used to accurately characterize EF. Thus, EPA only used aquatic sites with sufficient
data to calculate a reasonably reliable EF value.
To calculate the EF of aquatic systems, EPA searched its collection of selenium
concentration measurements from field studies (see Section 2.7.8 for a description of data
80

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sources and acceptability criteria) and identified aquatic sites with measurements from both
particulate material and the water column. EPA first identified all selenium measurements from
algae, detritus, or sediment, and then searched for corresponding water column measurements
from samples collected at the same aquatic site within one year of the particulate sample. If more
than one water measurement was available for any given particulate measurement, the median
was used. For each of these matched pairs of particulate and water measurements, EPA
calculated the ratio of particulate concentration to water concentration. If more than one ratio for
any given category of particulate material (algae, detritus, or sediment) was calculated at an
aquatic site, EPA used the median ratio. The geometric mean of the algae, detritus, and sediment
ratios was used as the site EF. Because there were at most only three possible values (one for
algae, one for detritus, and one for sediment), EPA used the geometric mean in order to reduce
the potential for one of the values to have excessive influence on the final site EF value.
The availability of selenium measurements from particulate material was limited. In
addition, the majority of particulate measurements were from sediment samples with a
significantly lower correlation to selenium in water (r = 0.34) compared to algae (r = 0.68; Fisher
r-to-z transformation, P < 0.001) and detritus (r = 0.94; Fisher r-to-z transformation, P < 0.001).
Therefore, to reduce uncertainty in estimating site-specific EF values, EPA limited its analysis to
those aquatic sites with at least two particulate selenium measurements with corresponding water
column measurements, and only used sediment measurements if there was at least one other
measurement from either algae or detritus. On the basis of these requirements, EF values were
calculated for 96 individual aquatic sites.
3.2.3 Food-Web Models
For the aquatic sites with a calculated EF value, EPA modeled the food webs for the fish
species the studies indicated were present. Some of those studies provided information about the
species and proportions of organisms ingested by fish, either through direct analysis of stomach
contents, or examination of the presence and prevalence of invertebrate species. For those
studies, that site-specific information in the food web models was used. Most studies, however,
did not provide site-specific food web information. In those cases, the food webs of fish species
present were modeled using information about their typical diet and/or eating habits obtained
from the NatureServe database (http://www.natureserve.org).
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After EPA developed food web models, EPA identified the appropriate species-specific
TTF values for each model and calculated	Although individual TTF values were
derived for several different taxa of invertebrates and fish (Table 3.10 and Table 3.11), some of
the food web models included one or more taxa for which no TTF value was available. EPA
estimated TTF values for these taxa using the same taxonomic approach used to estimate egg-
ovary to whole body, egg-ovary to muscle, and muscle to whole body conversion factors for taxa
without sufficient data. In brief, for taxa with insufficient data to calculate a TTF value, EPA
sequentially considered higher taxonomic classifications until one or more taxa for which a TTF
value was available matched the taxon being considered. If the lowest matching taxon was
common to more than one species with a TTF value available, EPA used the median TTF from
the matching species. EPA used site-specific food-web models to translate the egg-ovary
criterion element to a set of water column concentrations in order to derive the water column
concentration element of the selenium criterion. Details of these food web models are shown in
Table B-8 of Appendix B.
3.2.4 Classifying Categories of Aquatic Systems
Transformation reactions that convert dissolved selenium to particulate forms are the
primary route of entry into aquatic food webs, and are critical steps in selenium bioaccumulation
and toxicity (Chapman et al. 2010). Site-specific characteristics can result in substantial
bioaccumulation variability and consequently different risks of selenium toxicity for a given
dissolved selenium concentration. Freshwater systems fall into two distinct categories: lotic
systems such as rivers and streams, characterized by flowing water, and lentic systems, such as
lakes and ponds, characterized by largely standing water (e.g., Jones 1997). Water residence time
is generally shorter in lotic systems than in lentic systems, and subsequently, aquatic organisms
living in lentic systems tend to bioaccumulate more selenium than organisms living in lotic
systems for a given dissolved selenium concentration (ATSDR 2003; EPRI 2006; Luoma and
Rainbow 2005; Orr et al. 2006; Simmons and Wallschlagel 2005).
Although the distinction between lotic and lentic aquatic systems is often straightforward,
some aquatic systems possess both lotic and lentic characteristics. For example, flow rate can
vary greatly among lotic systems, with slow flowing low gradient systems (such as sloughs)
having longer residence times relative to fast flowing high gradient systems. Lotic systems can
also become more lentic during dry periods as hydrologic connectivity between deeper pools
82

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decrease or cease with decreasing flow (Buffagni et al. 2009). Downstream reaches of some low
gradient coastal rivers can also be influenced by tides (Riedel and Sanders 1998). Some lentic
systems can exhibit some degree of flow, such as lakes fed and drained by one or more streams;
however, the magnitude of flow is generally small compared to a lotic system. Even after
accounting for flow, the majority of water movement in a lentic system is driven typically by
wind or convection rather than gravity (e.g., Jones 1997). Finally, human-made reservoirs have
some features that are intermediate between typical lotic and lentic systems. For example,
reservoirs tend to be longer and narrower than natural lakes, and they have somewhat shorter
water retention time than a natural lake of comparable volume (Thornton et al. 1990). Overall,
however, reservoirs as a general class are considered more lentic than lotic, and have historically
been classified as a type of lake (Thornton et al. 1990).
To verify the suitability of lentic and lotic aquatic system categories as the basis for
independent water column criterion element values, EPA evaluated the aquatic systems that
provided data for the 96 EF values. EPA utilized the description provided by the study authors to
categorize each aquatic system as either lotic or lentic. Of the 39 lentic sites, the authors
identified them as ponds (n = 18), lakes (n = 13), reservoirs (n = 4), or marshes (n = 4). Of the 57
lotic sites, the authors identified them as creeks (n = 31), rivers (n = 16), artificial channels
(drains and wasteways, n = 3), springs (n = 2), sloughs (n = 2), or ephemeral systems (draws and
washes, n = 3). The three ephemeral aquatic sites (two washes and one draw) were categorized as
lotic because there was flowing water at the time they were sampled (Butler et al. 1995; Presser
and Luoma 2009). EF values for these aquatic systems are shown in Figure 3.6.
83

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OMa
8
Lake Reserv. Marsh Pond Creek Spring River Drain Ephem. Slough
Lentic
Lotic
Figure 3.6. Enrichment Factors (EF) for 96 Aquatic Sites Derived from Published Studies
and Organized by Waterbody Type.
The dashed line represents the median EF for the 39 lentic sites (0.9 L/g), and the solid line
represents the median EF for the 57 lotic sites (0.4 L/g). See text for information on labeled data
points.
Because the six labeled aquatic sites in Figure 3.6 (Ma, Ba, Bn, Hi, El and Fl) appear as
outliers, EPA selected them for further scrutiny. Data from site "Ma" result in an EF value of 5.2
L/g. Site "Ma" was a small irrigation pond within the Mancos River Valley watershed in
southwestern Colorado (Butler et al. 1997). This watershed drains the Mancos Shale, a region
that is naturally high in selenium. Data from sites "Hi," "Bn," and "Ba" resulted in EF values of
5.0, 5.9, and 12.5 L/g, respectively. Data from site "Hi" were from High Rock Lake, NC, data
from site "Bn" were from Barnes Lake, British Columbia (Orr et al. 2006), and data from site
"Ba" was from Badin Lake, NC (Lemly 1985). The high EF values at these three lakes were the
result of a relatively high selenium concentration in particulate matter coupled with low aqueous
84

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selenium concentrations. Data from site "El" result in an EF value of 6.3 L/g. Site "El" is an
upstream site in the Elk River watershed in southeastern British Columbia, and the relatively
large EF is the primarily the result of a low aqueous selenium concentration (McDonald and
Strosher 1998). Data from "Fl" result in an EF value of 7.1 L/g. Site "Fl" is within Flathead
wetland in southeastern British Columbia, and the relatively large EF is primarily the result of a
low aqueous selenium concentration (Orr et al. 2012).
Figure 3.7 illustrates the variability in EF values across aquatic systems and substantial
overlap between lotic and lentic categories. Some of this variability can be attributed to
differences in the ambient concentration of selenium in the water column at these aquatic sites.
EF is the ratio of selenium in particulate material (Cparticuiate) to selenium in the water column
(Cwater)- As expected, the selenium concentrations in particulate material are positively correlated
with the selenium concentrations in the water column (Figure 3.7A). The plot of Cparticuiate versus
Cwater shows a significant (P<0.001) positive relationship for both lentic (slope = 0.65, 95%
confidence interval = [0.50, 0.80]) and lotic (slope = 0.55, 95% confidence interval = [0.43,
0.68]) aquatic systems. However, selenium accrual in particulate matter is lower at aquatic sites
with a higher water concentration of selenium compared to aquatic sites with a lower water
concentration of selenium (Figure 3.7B). The plot of Cwater versus EF shows a significant
(P0.001) negative relationship for both lentic (slope = -0.36, 95% confidence interval = [-0.51, -
0.22]) and lotic (slope = -0.42, 95% confidence interval = [-0.55, -0.30]) aquatic systems.
Consistent with other studies (e.g., Hamilton and Palace 2001; Brix et al. 2005; Orr et al. 2006),
these results illustrate that the overall longer residence times of lentic systems result in increased
bioaccumulation of selenium compared to lotic systems.
85

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100
10.00
Lentic R2 = 0.68
Lotic R2 = 0.59
*
1.00
Particulate
(Hg/g)
X )tjn\ *
X
EF (L/g)
0.10
* Lotic
* Lotic
Lentic R2 = 0.42
Lentic
 Lentic
Lotic R2 = 0.46
0.01
100
10.0
100.0
Water (|xg/L)
Water (fig/L)
Figure 3.7. The Relationship between Cwater and Cparticulate, and Cwater and EF for the 39
Lentic and 57 Lotic Aquatic Systems.
A: Relationship between Cwater and Cparticuiate by site category.
B: Relationship between Cwater and Eh' by site category.
Solid line, ordinary least squares linear regression of logged data from lentic aquatic systems.
Dashed line, ordinary least squares linear regression of logged data from lotic aquatic systems.
Figure 3.8 shows the distribution of EF values grouped by lotic and lentic aquatic system
categories. Although EPA derived the lentic and lotic EF values from aquatic sites with a similar
range of water concentrations, the relative proportion of EF values collected at sites with higher
water concentrations is larger for lentic sites than lotic sites. In particular, 6 of the 39 lentic EF
values were from ponds in the Kesterson National Wildlife Refuge where Cwater ranged from
38.6-196 |ig/L (Saiki and Lowe 1987; Schuler et al. 1990). Despite the influence of selenium
water concentration on EF, the median of EF values from lentic and lotic aquatic systems are
significantly different from each other (Mann-Whitney U, P = 0.002). EPA concludes from these
analyses that lentic and lotic aquatic system categories are appropriate categories for
differentiating Se bioaccumulation. A listing of all aquatic-sites from which EFs were calculated
is provided in Appendix H.
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8
6
O)
LD
4
2
0
_Q_
Lentic (n=39)
Lotic (n=57)
Waterbody Type
Figure 3.8. Distribution of EF Values for the Same 96 Aquatic Systems.
(As shown in Figure 3.6 and Figure 3.7 grouped by lentic and lotic aquatic system categories).
Boxes show the 25th centile, median, and 75th centile EF values; whiskers show the 10th and 90th
centiles. Circles represent EF values greater than 1.5 times the interquartile (25th-50th - lower
circles; 50th-75th -upper circles) range. Dashed line represents the median EF of all 96 sites (0.63
L/g). The EF value of 12.48 L/g from Badin Lake (Lemly 1985) is off scale.
3.2.5 Deriving Protective Water Column Concentrations for Lentic and Lotic System
Categories
To derive the water column element of the selenium criterion, EPA translated the egg-
ovary criterion element to a distribution of water column concentration values for lentic and lotic
aquatic systems. EPA uses the EF values calculated for 96 aquatic sites, available information
about the fish species present at those sites, and food web models of those fish species. Because
translation of the egg-ovary criterion element is site- and species-specific, several studies
identifying more than one species of fish could potentially provide more than one translated
water column concentration (one translated water value for each species). EPA considered using
87

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all water column values for all species present to generate distributions of translated water
column values from lentic and lotic aquatic sites. However, the number of reported fish species
at aquatic sites with an EF value varied from one to six fish species. Furthermore, the studies
providing data for 31 of the 96 sites with EF values do not provide information on the species of
fish that may have been present at the aquatic site. Because the number of fish species at an
aquatic site was not consistently reported, and because the number of reported fish species does
not necessarily indicate the number of species present at a site, EPA calculated one translated
egg-ovary criterion element to water column value for each aquatic site with both an EF value
and at least one reported fish species. When more than one species was reported at a site, the
EPA used the lowest translated water value for that site. Using this methodology, EPA translated
the egg-ovary FCV into water column concentrations at 26 lentic and 39 lotic aquatic sites. EPA
used these distributions of water concentration values translated from the egg-ovary criterion
element to derive chronic water column criterion element values for lentic and lotic aquatic
systems. Table 3.13 shows the model parameter values used to translate the egg-ovary criterion
element to site-specific water concentrations, and Figure 3.9 shows the distribution of the
translated values.
88

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Table 3.13. Data for the 65 Site Minimum Translations of the Egg-Ovary Criterion Concentration Element to a Water Column
Concentration.
lik'iiliriciilion
Model Pnrnmeleis
Tmnsliilion
UcTcmuv
Silo
Species
1 > |)0
/:r
( / *'

(
x U.lIlT
1! irk nor 1 TX
Lasi Allen Rosen on; Mcdioinc Bow \VY
low a daricr
Lcnlic
2.} 1
1.45
"" X"
1.5"
Birkner 1978
Galett Lake, Laramie WY
Iowa darter
Lentic
0.88
1.45
2.87
4.15
Birkner 1978
Larimer Highway 9 Pond, Fort Collins CO
northern plains killifish
Lentic
1.70
1.20
2.44
3.04
Birkner 1978
Meeboer Lake, Laramie WY
northern plains killifish
Lentic
0.58
1.20
2.44
8.96
Birkner 1978
Miller's Lake, Wellington CO
Iowa darter
Lentic
2.37
1.45
2.87
1.53
Birkner 1978
Sweitzer Lake, Delta CO
fathead minnow
Lentic
0.87
1.40
2.78
4.45
Birkner 1978
Twin Buttes Reservoir, Laramie WY
Iowa darter
Lentic
1.21
1.45
2.87
3.01
Bowie et al. 1996
Hyco Reservoir
bluegill
Lentic
2.35
2.13
2.00
1.51
Butler etal. 1993
Navajo Reservoir, Piedra River Arm, near
La Boca
brown trout
Lentic
1.26
1.45
2.78
2.98
Butler etal. 1997
Large pond south of G Road, southern
Mancos Valley
fathead minnow
Lentic
2.00
1.40
2.78
1.94
Butler etal. 1997
Pond downstream from site MNP2, southern
Mancos Valley
smallmouth bass
Lentic
5.15
1.42
1.93
1.07
Butler etal. 1997
Pond on Woods Canyon at 15 Road
fathead minnow
Lentic
0.90
1.40
2.78
4.29
Grasso etal. 1995
Arapahoe Wetlands Pond
fathead minnow
Lentic
0.86
1.40
2.78
4.49
Lemly 1985
Badin Lake
red shiner
Lentic
12.48
1.95
2.27
0.27
Lemly 1985
Belews Lake
red shiner
Lentic
1.75
1.95
2.27
1.94
Lemly 1985
High Rock Lake
red shiner
Lentic
4.99
1.95
2.27
0.68
Muscatello and
Janz 2009
Vulture Lake
northern pike
Lentic
1.01
2.39
4.02
1.56
Orr et al. 2012
Clode Pond 11
cutthroat trout
Lentic
0.71
1.96
2.29
4.70
Orretal. 2012
Elk Lakes 14
cutthroat trout
Lentic
1.64
1.96
2.29
2.05
Orr etal. 2012
Fording River Oxbow 10
cutthroat trout
Lentic
1.34
1.96
2.29
2.50
Orretal. 2012
Henretta Lake 27
cutthroat trout
Lentic
0.50
1.96
2.29
6.72
Saiki and Lowe
1987
Kesterson Pond 11
western mosquitofish
Lentic
0.51
1.20
2.37
10.52
Saiki and Lowe
1987
Kesterson Pond 2
western mosquitofish
Lentic
0.32
1.20
2.37
16.83
Saiki and Lowe
1987
Kesterson Pond 8
western mosquitofish
Lentic
0.60
1.20
2.37
8.84
Saiki and Lowe
1987
Volta Pond 26
western mosquitofish
Lentic
0.93
1.20
2.37
5.69
89

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lik'iiliriciilion
Model Pnrnmeleis
Tmnsliilion
UcTcmuv
Silo
Species
1 > |)0
/:r
( / *'

(
x U.lIlT
Stephens et al.
1988
Marsh 4720
common carp
Lentic
0.10
1.92
1.58
52.02
Butler etal. 1991
Uncompahgre River at Colona
rainbow trout
Lotic
0.63
2.44
2.33
4.21
Butler etal. 1993
Spring Cr. at La Boca
brown trout
Lotic
0.18
1.45
2.78
20.97
Butler etal. 1995
Hartman Draw near mouth, at Cortez
fathead minnow
Lotic
0.15
1.40
2.78
26.04
Butler etal. 1995
McElmo Cr. at Hwy. 160, near Cortez
fathead minnow
Lotic
0.90
1.40
2.78
4.32
Butler etal. 1995
McElmo Cr. downstream from Alkali Cyn.
fathead minnow
Lotic
0.37
1.40
2.78
10.57
Butler etal. 1995
McElmo Cr. downstream from Yellow
Jacket Cyn.
red shiner
Lotic
0.12
1.95
2.27
28.34
Butler etal. 1995
McElmo Cr.upstream from Yellow Jacket
Cyn.
red shiner
Lotic
0.10
1.95
2.27
35.60
Butler etal. 1995
Navajo Wash near Towaoc
speckled dace
Lotic
0.20
1.95
1.36
29.07
Butler etal. 1995
San Juan River at Four Comers
red shiner
Lotic
0.26
1.95
2.27
12.97
Butler etal. 1995
San Juan River at Mexican Hat Utah
common carp
Lotic
0.29
1.92
1.58
17.24
Butler etal. 1995
Woods Cyn. Near Yellow Jacket
fathead minnow
Lotic
0.40
1.40
2.78
9.60
Butler etal. 1997
Cahone Canyon at Highway 666
green sunfish
Lotic
0.20
1.45
2.29
23.22
Butler etal. 1997
Mud Creek at Highway 32, near Cortez
fathead minnow
Lotic
0.07
1.40
2.78
55.27
Casey 2005
Deerlick Creek
rainbow trout
Lotic
2.24
2.44
2.33
1.18
Casey 2005
Luscar Creek
rainbow trout
Lotic
0.33
2.44
2.33
8.14
Formation 2012
Crow Creek - 1A
brown trout
Lotic
0.80
1.45
2.96
4.42
Formation 2012
Crow Creek - 3 A
brown trout
Lotic
0.81
1.45
2.97
4.37
Formation 2012
Crow Creek - CC150
brown trout
Lotic
1.04
1.45
2.91
3.44
Formation 2012
Crow Creek - CC350
brown trout
Lotic
1.16
1.45
2.97
3.02
Formation 2012
Crow Creek - CC75
brown trout
Lotic
1.19
1.45
2.87
3.07
Formation 2012
Deer Creek
brown trout
Lotic
1.55
1.45
3.00
2.25
Formation 2012
Hoopes Spring - HS
brown trout
Lotic
0.24
1.45
3.86
11.06
Formation 2012
Hoopes Spring - HS3
brown trout
Lotic
0.54
1.45
2.63
7.40
Formation 2012
Sage Creek - LSV2C
brown trout
Lotic
0.45
1.45
3.01
7.76
Formation 2012
Sage Creek - LSV4
brown trout
Lotic
0.69
1.45
2.88
5.22
Formation 2012
South Fork Tincup Cr.
brown trout
Lotic
1.32
1.45
3.05
2.58
Hamilton and
Buhl 2004
Lower East Mill Creek
cutthroat trout
Lotic
1.32
1.96
2.29
2.55
McDonald and
Strosher 1998
Elk R. above Cadorna Cr. (745)
mountain whitefish
Lotic
6.30
7.39
2.97
0.11
McDonald and
Strosher 1998
Fording R. above Swift Cr. (746)
cutthroat trout
Lotic
0.23
1.96
2.29
14.91
Orretal. 2012
Elk River 1
cutthroat trout
Lotic
0.55
1.96
2.29
6.14
90

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lik'iiliriciilion
Model Pnrnmeleis
Tmnsliilion
UcTcmuv
Silo
Species
1 > |)0
/:r
( / *'

(
x U.lIlT
Orretal. 2012
Elk River 12
cutthroat trout
Lotic
2.67
1.96
2.29
1.26
Orretal. 2012
Fording River 23
cutthroat trout
Lotic
0.21
1.96
2.29
16.20
Orretal. 2012
Michel Creek 2
cutthroat trout
Lotic
0.28
1.96
2.29
11.85
Saiki and Lowe
1987
San Luis Drain
western mosquitofish
Lotic
0.36
1.20
2.37
14.81
Saiki and Lowe
1987
Volta Waste way
western mosquitofish
Lotic
1.03
1.20
2.37
5.17
Saiki etal. 1993
Mud Slough at Gun Club Road
bluegill
Lotic
1.37
2.13
1.47
3.53
Saiki etal. 1993
Salt Slough at the San Luis National
Wildlife Refuge
bluegill
Lotic
0.43
2.13
1.47
11.29
Saiki etal. 1993
San Joaquin R. above Hills Ferry Road
bluegill
Lotic
0.36
2.13
1.47
13.50
Saiki etal. 1993
San Joaquin R. at Durham Ferry State
Recereation Area
bluegill
Lotic
0.75
2.13
1.47
6.46
a - Geometric mean of the median enrichments functions (/.'/<) for all available food types (algae, detritus, and sediment). EF (L/g) = Cf00d/Cwater-
b - Taxa-specific conversion whole-body to egg ovary conversion factor (CF; dimensionless ratio),
c - Composite trophic transfer factor (777'"'""v"'v'"'). Product of TTF values for all trophic levels.
d - Translated water concentration corresponding to an egg-ovary criterion element of 15.1 mg Se/kg dw, calculated by Equation 18.
91

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Cumulative
proportion
~
m
I x *

!  *

i i *x

 *
  X

 *
.  X

m X
X

	X
	*
"  X

1 *
J *
* X

t X
' xx

 , X
 X * Lotic

,  | ;
*  Lentic
i 1

X .
* I
 X
X I
m *
X *
1 1 1 ( ( ( ( 1 |
 . 		
0
1
10
100
Water (n-g/'L)
Figure 3.9. Probability Distribution of the Water Column Concentrations Translated from
the Egg-Ovary Criterion Element at 26 Lentic and 39 Lotic Aquatic Sites.
Dashed and dash-dot lines show the 20th percentiles of the lentic and lotic distributions,
respectively.
EPA selected the 20th percentile from the distribution of translated water column values
of each category as the final national water column criterion element concentrations (3.1 |ig/L
for lotic waters and 1.5 |ig/L for lentic waters) because the 20th percentile is consistent with past
practice as it provides a high probability of protection for most aquatic systems in both lentic and
lotic categories. Table 3.14 provides the 20th percentile of the water concentration values
translated from the egg-ovary criterion element value.
Table 3.14. Water Column Criterion Element Concentration Values Translated from the
Egg-Ovary Criterion Element.

l.entic
Lotic
20th percentile
(final EPA-recommended water column criterion element)
1.5 jig/L
3.1 jig/L
As discussed in Section 2.2.2, selenium bioaccumulation potential depends on several
biogeochemical factors that characterize a particular aquatic system. Uncertainty in the
translation of the egg-ovary criterion element to the water column element can be reduced by
deriving a site-specific criterion that uses site-specific selenium data and information on food-
web dynamics from a biological assessment of the aquatic system. The general considerations
are provided in Appendix K. The derivation of water column criterion element values described
92

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above is constrained by the need to apply a national criterion value to a large number of aquatic
systems for lentic and lotic systems.
3.2.6 Derivation of Averaging Period for Chronic Water Criterion Element
In the context of selenium bioaccumulation in a single trophic level, k would be the first-
order depuration coefficient, and 1/k would equal the time needed to depurate to a concentration
of 1/e times the initial concentration (where e=2.718). For depuration of two trophic levels
sequentially, invertebrates and fish, the characteristic time is likewise the time needed for c/c0 to
reach a value of 1/e. This differs from typical criteria averaging periods based on U.S. EPA
(1995), where the concept that the criterion averaging period should be less than or equal to the
"characteristic time" describing the toxic speed of action due to direct waterborne toxicity of
metals (i.e., where characteristic time = 1/k, where k is the first-order kinetic coefficient in a
toxicokinetic model fitted to the relationship between LC50 and exposure duration). For the first
trophic level, the kinetics for algal bioaccumulation and depuration were assumed to be rapid
compared to the kinetics for larger organisms at higher trophic levels; that is, the characteristic
time for algae was assumed to be negligible.
For the second trophic level, invertebrates, values for kTL2 are tabulated elsewhere in the
document. A value of 0.1/day appears to be environmentally conservative, considerably higher
than those for Lumbriculus, Asian clam, and zebra mussel, but slightly lower than copepods,
which are very small in size. This corresponds to a characteristic time of 10 days.
For fish, the median depuration coefficient measured by Bertram and Brooks (1986) for
6-9 month-old (early adult) fathead minnows is applied, providing a kTL3 value of 0.02/day. This
corresponds to a characteristic time of 50 days. Because of the small size of adults of this
species, this represents faster kinetics than would likely be applicable to the salmonids and
centrarchids of greatest concern for selenium toxicity, consonant with the Newman and Mitz
(1988) inverse relationship between depuration rate and organism size. The striped bass k value
of Baines et al. (2002) is inapplicable here because it was measured in the early juvenile life
stage, a size that is too small to be relevant to reproductive impairment stemming from exposure
of adult females.
As shown in Appendix J, the characteristic time for the combined second and third
trophic levels (invertebrates and fish) is the approximate sum of the above two characteristic
times, or 60 days. The analysis of the protectiveness of a 30-day averaging period, shorter than
93

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the characteristic time, was performed and is shown in Appendix J. That analysis demonstrated
that a 30-day averaging period for the chronic water criterion element affords protection under
all conditions, and is therefore the duration recommended for the chronic water column criterion
element.
3.3 Intermittent-Exposure Water Criterion Element: Derivation from
the Chronic Water Criterion Element
Chapman et al. (2009) noted that selenium acute toxicity has been reported rarely in the
aquatic environment and that traditional methods for predicting effects based on direct exposure
to dissolved concentrations do not work well for selenium. As demonstrated in Appendix J, the
kinetics of selenium accumulation and depuration are sufficiently slow that attainment of the
water criterion element concentration by ambient 30-day averages will protect sensitive aquatic
life species even where concentrations exhibit a high degree of variability.
To address situations where pulsed exposures of selenium could result in
bioaccumulation in the ecosystem and potential chronic effects in fish, EPA is providing an
intermittent exposure water criterion element concentration intended to limit cumulative
exposure to selenium, derived from the chronic 30-day water criterion element magnitude and
from its duration, which was obtained from the kinetic analysis of Appendix J. That is, the
intermittent criterion element is based on the same kinetic analysis used to derive the water
chronic averaging period (30 days).
To illustrate the concept of the intermittent criterion element and its dependence on the
30-day criterion element magnitude and duration, Figure 3.10 shows a possible sequence of
exposures over a 30-day period.
94

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Figure 3.10. Illustration of Intermittent Spike Exposure Occurring for a Certain
Percentage of Time (e.g., 10%) Over a 30-Day Period, and Background Exposure
Occurring for the Remaining Percentage of Time (e.g., 90%).
The 30-day average concentration, C30day, is given by Equation 20:
C30-day ~ Cintfint Cbkgmd(1  f int)
(Equation 20)
Where:
Qnt = the intermittent spike concentration (|ig/L)
fint = the fraction of any 30-day period during which elevated selenium
concentrations occur
Cbkgmd = the average daily background concentration occurring during the
remaining time, integrated over 30 days.
C30-day i s not to exceed the chronic criterion element, WQC3o-day. If the intent is to apply a
criterion element, WQCmt, to the intermittent spike concentrations, then replacing C,nt with
WQCmt and Cso-day with WQC3o-day in the above equation, and then solving for WQCmt yields
Equation 21:
WQCint =
WQC^q day Cbkgrndi.1 fint)
fint
(Equation 21}
95

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The equation expresses the intermittent exposure water criterion element in terms of the
30-day average chronic water criterion element, for a lentic or lotic system, as appropriate, while
accounting for the fraction in days of any 30-day period the intermittent spikes occur and for the
concentration background occurring during the remaining time. The reasonable worst-case
assumption inherent in this approach is that selenium bioaccumulation is linear over a very wide
range of concentrations, that is, EF and TTF values do not decrease significantly as
concentrations increase.
If the heights of three spikes in Figure 3.10 were to differ somewhat among each other,
the intermittent element would apply to the arithmetic mean of the three. If the background
concentrations were to vary somewhat, then the arithmetic mean background would be entered
into the equation. Where concentrations vary smoothly over time, it does not matter where the
line is drawn defining elevated versus background concentrations. The intermittent element will
yield the same level of protection as the 30-day average element, provided that the equation uses
(a) the average of the concentrations occurring for the fraction of time defined as being
intermittently elevated, and (b) the average of the concentrations occurring for the remaining
time, defined as being background. The intermittent element will only be exceeded under
conditions that would have caused the 30-day element to be exceeded, had it been applied.
Table 3.15 illustrates example values for the intermittent element. The bottom row of the
lotic and lentic values and the right column are to emphasize that WQCmt '\ not an independent
element but a re-expression of the 30-day average water criterion element concentration. WQCmt
converges to WQCw-cby when the background concentration is already at WQC3o-day or when the
intermittent exposure is said to occur throughout the 30-day period.
96

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Table 3.15. Representative Values of the Intermittent Water Criterion Element
Concentration.
likurml


l-'raction oi l
ime. in a 30-dav period

Cone.
0.03333
0.05
0.1
0.2
0.5
1
^ hlt^niil
(1 day)
(1.5 (lavs)
(3 (lavs)
(6 davs)
(15 davs)
(30 davs)
(MK/I-)
l.olic Inlerniillenl Criterion Klcment. II OC,m(iig/l.)
U
93
t>2
31
15.5
t>.2
3.1
1
64
43
22
11.5
5.2
3.1
2
35
24
13
7.5
4.2
3.1
2.5
20.5
14.5
8.5
5.5
3.7
3.1
3 1
3 1
3 1
3 1
3 1
3 1
3 1

l.enlic Inlerniillenl Criterion Klemenl. H (X';,,,(ii/l.)
0
45
30
15
7.5
3
1.5
0.5
30.5
20.5
10.5
5.5
2.5
1.5
1
16
11
6
3.5
2
1.5
1.25
8.8
6.3
3.8
2.5
1.8
1.5
1.5
1.5
1.5
1.5
1.5
1.5
1.5
If the value of fnt, the intermittent exposure fraction of the month, is assigned a value less
than one day, the intermittent criterion element value could exceed water concentrations that
have been shown to be acutely toxic to sensitive species in 2- or 4-day toxicity tests (compiled in
U.S. EPA 2004). Because the concentrations that would be acutely toxic in exposures of less
than one day might not be much greater than those observed to be toxic in 2-4 day exposures, the
intermittent fraction of the month must not be assigned a value less than 0.033, corresponding to
one day.
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4 National Criterion for Selenium in Fresh Waters
The available data indicate that freshwater aquatic life would be protected from the toxic
effects of selenium by applying the following four-part criterion, recognizing that fish tissue
elements supersede the water elements (except in special situations, see footnotes 3 and 4, Table
4.1) and that the egg-ovary tissue element supersedes all other tissue elements:
1.	The concentration of selenium in the eggs or ovaries of fish does not exceed 15.1 mg/kg,
dry weight; 1
2.	The concentration of selenium (a) in whole-body of fish does not exceed 8.5 mg/kg dry
weight, or (b) in muscle tissue of fish (skinless, boneless fillet) does not exceed 11.3
mg/kg dry weight;
3.	The 30-day average concentration of selenium in water does not exceed 3.1 |ig/L in lotic
(flowing) waters and 1.5 |ig/L in lentic (standing) waters more than once in three years
on average;
4.	The intermittent concentration of selenium in either a lentic or lotic water, as appropriate,
i,i tAmr	WQC30-(iay ~ Cbkgrnd(l~f int)	,1	 ,1
does not exceed WQCint = 			-	more than once in three years on
/ int
average.3
98

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Table 4.1. Summary of the Recommended Freshwater Selenium Ambient Chronic Water
Quality Criterion for Protection of Aquatic Life.	
Media
Type
Criterion
Klcmcnl
Magnitude
Duration
l-'requencv
I'isli Tissue1
K jig/Ovary 2
I'isli W hole
liody or
Muscle'
\Yater Column4
Monthly
Average
Kxposurc
1 n 1 erin i11ent K x pos u re"
1 5 I mg/kg dw
8.5 mg/kg dw
whole body
or
11.3 mg/kg
dw muscle
(skinless,
boneless filet)
1.5 |ig/L in
lentic aquatic
systems
3.1 |ig/L in lotic
aquatic systems
WQCint =
WQCzo-day ~ Cftkgrnd^X ~ f int)
fint
Instantaneous
measurement6
Instantaneous
measurement6
30 days
Number of days/month with an
elevated concentration
Not to be
exceeded
Not to be
exceeded
Not more than
once in three
years on
average
Not more than once in three years
on average
1.	Fish tissue elements are expressed as steady-state.
2.	Egg/Ovary supersedes any whole-body, muscle, or water column element when fish egg/ovary concentrations are
measured, except as noted in footnote 4 below.
3.	Fish whole-body or muscle tissue supersedes water column element when both fish tissue and water concentrations are
measured, except as noted in footnote 4 below.
4.	Water column values are based on dissolved total selenium in water and are derived from fish tissue values via
bioaccumulation modeling. When selenium inputs are increasing, water column values are the applicable criterion element
in the absence of steady-state condition fish tissue data.
5.	Where WQC30-day is the water column monthly element, for either a lentic or lotic waters; Cbkgrnd is the average
background selenium concentration, and fmt is the fraction of any 30-day period during which elevated selenium
concentrations occur, with fint assigned a value >0.033 (corresponding to 1 day).
6.	Fish tissue data provide instantaneous point measurements that reflect integrative accumulation of selenium over time and
space in fish populations) at a given site.
EPA recommends that states and tribes adopt into their water quality standards a
selenium criterion that includes all four elements, and express the four elements as a single
criterion composed of multiple parts, in a manner that explicitly affirms that the whole-body or
muscle elements supersede the water column element, and the egg-ovary element supersedes any
other element. The magnitude of the fish egg-ovary element is derived from analysis of the
99

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available toxicity data. The magnitudes of the fish whole-body element and fish muscle elements
are derived from the egg-ovary element coupled with data on concentration ratios among tissues.
The magnitudes of the water column elements are derived from the egg-ovary element coupled
with bioaccumulation considerations. Inclusion of the fish whole-body or fish muscle element
into the selenium criterion ensures the protection of aquatic life when fish egg or ovary tissue
measurements are not available, and inclusion of the water column elements into the selenium
criterion ensures protection when neither fish egg-ovary nor fish whole-body nor muscle tissue
measurements are available. To ensure that the contribution of short-term exposures to the
bioaccumulation risks is accounted for in all situations, EPA is also recommending that the
intermittent exposure element be included in the selenium criterion, as noted above. EPA is not
recommending a separate acute criterion derived from the results of toxicity tests having water-
only exposure because selenium is bioaccumulative and toxicity primarily occurs through dietary
exposure. Application of the intermittent exposure criterion element values to single day, high
exposure events will provide protection from the most important selenium toxicity effect,
reproductive toxicity, by protecting against selenium bioaccumulation in the aquatic ecosystem
resulting from short-term, high exposure events. It is unnecessary to have an additional acute
water column criterion element because the intermittent exposure criterion element will be more
stringent than an acute criterion element. Further, as noted in this document, there have been few
if any acute exposure, water column-only selenium aquatic toxicity events documented in the
literature.
In implementing the water quality criterion for selenium under the NPDES permits
program, states may need to establish additional procedures due to the unique components of the
selenium criterion. Where states use the selenium water column concentration criterion element
values only (as opposed to using both the water column and fish tissue elements) for conducting
reasonable potential (RP) determinations and establishing water quality-based effluent
limitations (WQBELS) per 40 CFR 122.44(d), existing implementation procedures used for
other acute and chronic aquatic life protection criteria may be appropriate. However, if states
also decide to use the selenium fish tissue criterion element values for NPDES permitting
purposes, additional state WQS implementation procedures (IPs) will be needed to determine the
need for and development of WQBELs necessary to ensure that the fish tissue criterion
element(s) are met.
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EPA recommends that states use the default monthly average exposure water column
elements of the criterion, adopted as part of the state's water quality criterion. Alternatively,
states may want to develop and adopt, and submit for EPA approval, either a site-specific water
column criterion (see Appendix K for details), or a procedure to facilitate the translation of a fish
tissue criterion element concentration into site-specific water concentration values. A site-
specific water column criterion element or set of lentic/lotic criterion element values may be
developed using a mechanistic modeling approach (Presser and Luoma 2010) or using the
empirical bioaccumulation factor approach, both described in Appendix K, for the specific
waterbody or waterbodies, or a on a state-wide basis. A translation procedure must be
scientifically defensible and able to produce repeatable and predictable outcomes, and must be
consistent with either the mechanistic modeling approach or the empirical bioaccumulation
factor approach described in Appendix K. The chronic selenium criterion is derived to be
protective of the entire aquatic community, including fish, amphibians, and invertebrates. Fish
are the most sensitive to selenium effects. Selenium water quality criterion elements based on
fish tissue (egg-ovary, whole body, and/or muscle) sample data override the criterion elements
based on water column selenium data due to the fact, noted above, that fish tissue concentrations
provide the most robust and direct information on potential selenium effects in fish. However,
because selenium concentrations in fish tissue are a result of selenium bioaccumulation via
dietary exposure, there are two specific circumstances where the fish tissue concentrations do not
fully represent potential effects on fish and the aquatic ecosystem: 1) in "fishless" waters, and 2)
in areas with new selenium inputs.
Fishless waters are defined as waters with insufficient instream habitat and/or flow to
support a population of any fish species on a continuing basis, or waters that once supported
populations of one or more fish species but no longer support fish (i.e., extirpation) due to
temporary or permanent changes in water quality (e.g., due to selenium pollution), flow or
instream habitat. Because of the inability to collect sufficient fish tissue to measure selenium
concentrations in fish in such waters, water column concentrations will best represent selenium
levels required to protect aquatic communities and downstream waters in such areas.
New inputs are defined as new activities resulting in selenium being released into a lentic
or lotic waterbody. New inputs will likely result in increased selenium in the food web, likely
resulting in increased bioaccumulation of selenium in fish over a period of time until the new or
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increased selenium release achieves a quasi-"steady state" balance within the food web. EPA
estimates that concentrations of selenium fish tissue will not represent a "steady state" for several
months in lotic systems, and longer time periods (e.g., two to three years) in lentic systems,
depending upon the hydrodynamics of a given system such as the location of the selenium input
related to the shape and internal circulation of the waterbody, particularly in reservoirs with
multiple riverine inputs, hydraulic residence time, and the particular food web. Estimates of
steady state under new or increased selenium input situations are expected to be site dependent,
so local information should be used to better refine these estimates for a particular waterbody.
Thus, EPA recommends that fish tissue concentration not override water column concentration
in these situations until these periods of time have passed in lotic and lentic systems,
respectively, or steady state conditions can be estimated.
4.1 Protection of Downstream Waters
EPA regulations at 40 CFR 131.10(b) provide that "[i]n designating uses of a waterbody
and the appropriate criteria for those uses, the state shall take into consideration the water quality
standards of downstream waters and ensure that its water quality standards provide for the
attainment and maintenance of the water quality standards of downstream waters." Especially in
cases where downstream waters are lentic waterbody types (e.g., lakes, impoundments), or
harbor more sensitive species, a selenium criterion more stringent than that required to protect
in-stream uses may be necessary to ensure that water quality standards provide for the attainment
and maintenance of the water quality standards of downstream waters.
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5 Site-specific Criteria
All four elements of the freshwater selenium criterion may be modified to reflect site-
specific conditions where the scientific evidence indicates that different values will be protective
of aquatic life and provide for the attainment of designated uses.
Since the fish egg-ovary criterion element is based on toxicity data, a state may modify
that element by applying the Recalculation Procedure (U.S. EPA 2013b) to edit the species
toxicity database to reflect taxonomic relatedness to the site assemblage, while including tested
surrogates for untested resident species.
It is important to note that species in the national data set that are not present at a site
should not be deleted from the data set because those species serve as surrogate(s) for other
species known or expected to be present at a site. Confidence in the applied tissue criterion
element can be improved by further testing of fish species resident at the site. The most relevant
testing would measure the survival and occurrence of deformities in offspring of wild-caught
female fish to determine an ECio for selenium in the eggs or ovaries (e.g., following Janz and
Muscatello 2008).
Using either the EPA national recommended egg-ovary, whole-body, or muscle criterion
concentration element or a site-specific egg-ovary, whole-body, or muscle criterion element,
translation of the fish tissue criterion to a protective water concentration can be performed in a
manner that accounts for site-specific conditions. Appendix K provides a step-wise process for
deriving each parameter used in Equation 18 to perform a site-specific translation. These steps
include:
1.	selecting a target fish species for the waterbody,
2.	determining the primary food source for the target species
3.	determining the appropriate TTF values,
4.	determining the appropriate EF value, and
5.	determining the appropriate CF value.
Appendix K also provides information on the data necessary to derive a site-specific
criterion, as well as scientifically defensible methods, including the use of traditional
Bioaccumulation Factors (BAFs), in addition to the more comprehensive mechanistic modeling.
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6 Effects Characterization
6.1 Fish and Amphibians
6.1.1 Principles for Using Studies for which ECms Cannot Be Calculated
When the data from an acceptable chronic test met the conditions for logistic regression
analysis, the ECio was used. When data did not allow calculation of ECs but did allow
calculation of closely spaced NOECs and LOECs, the NOEC was used to approximate the ECio.
No NOEC values were used in calculating the tissue criterion element values.
When significant effects were observed at all treatment concentrations, such that no
treatment concentration was classified as a NOEC, then the chronic value was assigned as "less
than" (<) the lowest tested concentration. When no significant effects were observed at any
concentration, such that no treatment concentration was defined as an LOEC, then the chronic
value was assigned as "greater than" (>) the highest tested concentration.
A number of the chronic values in Sections 3.1.1 and 6.1.2 (reproductive effects) and in
Section 6.1.9 (nonreproductive effects) include a "greater than" (>) or "less than" (<) sign
because of an inability to resolve an exact value when all exposure concentrations of a study
yielded either too little or too much effect to provide a point estimate of a chronic value. The
decision to use chronic values with a "greater than" or "less than" sign in calculating an SMCV
followed a rule based on whether these values add relevant information to the mean, as described
below. None of these values were used in this assessment to derive the tissue criterion element
values.
6.1.1.1 Evaluation Approach
	Neither a low "greater than" value nor a high "less than" value were used to
calculate the SMCV;
	Both a low "less than" value and a high "greater than" value were included in the
SMCV calculation. However, none of these values were used in this assessment to
calculate the numeric criterion values for fish tissue.
For example, a chronic value reported here as ">15 mg Se/kg" is ignored if the tentative
SMCV is 20 mg Se/kg. The ">15 mg Se/kg" value indicates that no significant effects were
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observed at the study's highest tested concentration of 15 mg Se/kg. As this is consistent with
what would be expected if the SMCV were 20 mg Se/kg, it provides no information to support
modifying the SMCV. However, a different study showing no effects at its highest tested
concentration and yielding the value ">25 mg Se/kg" is not consistent with an SMCV of 20 mg
Se/kg, and indicates that the ">25 mg Se/kg" value provides information for modifying the mean
upwards. Conversely, a chronic value reported here as "<15 mg Se/kg" indicates that significant
effects were observed even at the study's lowest tested concentration of 15 mg Se/kg. As this is
not consistent with a 20 mg Se/kg SMCV, it indicates the utility of the "<15 mg Se/kg"
information for modifying the SMCV downwards. On the other hand, a value reported here as
"<25 mg Se/kg" would not be used to recalculate a 20 mg Se/kg SMCV. The intent of the
approach is to use all quality information that is relevant and appropriate for calculating the
SMC Vs.
6.1.2 Acceptable Studies of Fish Reproductive Effects of Genera that were not among the Four
Most Sensitive Genera
The following is a brief synopsis of the experimental design, test duration, relevant test
endpoints, and other critical information regarding the genera that were not the four most
sensitive but were included in the number of GMCVs in the dataset (see Section 3.1.3), The
studies in this section involve effects on the offspring of exposed female fish. Data are
summarized in Table 3.1. Details of these studies are contained in Appendix C.
6.1.2.1 Cyprinidae
Pimephales promelas (fathead minnow)
Schultz and Hermanutz (1990) examined the effects of selenium transferred from
parental fish (females) on fathead minnow larvae. The parental fathead minnows were first
exposed to selenite (10 |ig/L) that was added directly to the water in artificial streams in a
mesocosm study. The selenite entered the food web and contributed to exposure via diet.
Spawning platforms were submerged into treated and control streams. Embryos were collected
from the spawning platforms and transferred to a proportional diluter where they were reared in
incubation cups for observation. Treated embryos in the egg cups were exposed to 10 |ig/L
selenium. Edema and lordosis were observed in approximately 25 percent of the larvae spawned
and reared in natural water spiked with 10 jag Se/L and in < 6 percent of control larvae. Although
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a case can be made that the selenium treatment had a higher rate of edema and lordosis, there are
some issues that add uncertainty to the estimation of an effect concentration (R. Erickson,
personal communication). Heavy mortality/loss of embryo/larvae during monitoring and the
erratic occurrence of the abnormalities (e.g., significant incidence of edema in only 3 of 10
replicates for the Se treatment) led to the conclusion that results should not be used for criterion
derivation. The data from this study support the range of reproductive effect levels determined in
other fish studies. The Se concentration in embryos from the 10 |ig/L treatment stream of 3.91
mg/kg ww converts to 25.6 mg/kg dw using 15.3% dw (N=3 range 14.7 - 15.6%) for fathead
minnow eggs (R. Erickson, personal communication).
Two other studies suggest fathead minnows are less sensitive to selenium than other fish.
Young et al. (2010) observed that fathead minnow populations remained after selenium
contamination of Belews Lake had eliminated most other fish species, including bluegill and
largemouth bass. In a maternal transfer laboratory study with fathead minnows, GEI (2008)
estimated ECioS for larval survival and deformities that ranged from 35 -65 mg Se/kg dw
expressed as maternal whole body, as noted in Appendix E, Figures E-2 and E-3.
6.1.2.2 Esocidae
Esox lucius (northern pike)
Muscatello et al. (2006) collected spawning northern pike from four sites near a uranium
milling operation in north-central Saskatoon, Canada, with egg concentrations ranging from 2.7
to 48 mg Se/kg dw. The four sites included a reference site and three sites 2, 10 and 15 km
downstream of the effluent discharge, representing a gradient of selenium exposure. Milt and ova
were stripped from gravid fish. Eggs were then fertilized in the field and incubated in the
laboratory for observations and measurements. The test was terminated when the majority of the
fry exhibited swim-up and had absorbed the yolk.
Mean egg diameter, fertilization success and cumulative embryo mortality were not
significantly different among the sites. Significant increases in percent total deformities
including edema, skeletal deformities, craniofacial deformities and fin deformities were observed
in fry originating from pike collected at the medium exposure site. The concentrations of
selenium in the northern pike eggs collected at the reference and low exposure site were very
similar, as were the percent total deformities in embryos/fry. The geometric mean of selenium in
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the eggs of the adult females at the reference and low exposure sites was 3.462 mg Se/kg dw and
the corresponding arithmetic mean of the percent total deformities was 13.20%. There were only
4 adult females from exposed sites, and all had relatively similar concentrations in their eggs, all
close to the geometric mean concentration of 34.00 mg Se/kg dw. Likewise, all four exposed
females had relatively similar percent total deformities, not far from their arithmetic mean of
33.40%. This is not a sufficient level of effect for applying TRAP to determine an ECio.
Furthermore, the relatively large spread between the two clusters of exposure concentrations
(3.462 and 34.00 mg Se/kg dw) would render a NOEC and LOEC unreliable and unsuitable for
defining a threshold. That is, the NOEC and LOEC would be "greater than" and "less than"
values, >3.462 and <34.00 mg Se/kg dw respectively, providing little information on the
sensitivity of northern pike compared to other species.
Instead, making use of the clustering of data at low exposure and effects and at elevated
exposure and effects, the effect level for the elevated exposure eggs was normalized to the low
exposure condition and rescaled to a 0-100% range. The rescaled (i.e., Abbott-adjusted) percent
of total deformities for the elevated exposure eggs was 24% (relative to the low exposure eggs).
Thus the concentration of selenium in the elevated exposure eggs (34 mg Se/kg dw) was
equivalent to an EC24, and is the only effects concentration that can be calculatedfor this test,
given the limitations in the range of concentrations tested and effects observed. Although the
EC24 is not directly translatable to an EC10 for use in determining the criterion, it is useful for
comparison with the EC24 in other species in order to determine species sensitivity rank. The
EC24 for skeletal deformities from the Holm et al. (2005) study of rainbow trout, calculated to be
30.9 mg Se/kg dw in eggs, is slightly lower than the northern pike value, indicating these two
species may be similar in tolerance, with the northern pike being slightly more tolerant (see
Appendix C for more details.)
6.1.2.3 Salmonidae
Seven publications provide quantitative data on the effects of selenium on salmonid
embryo/larval survival and deformity that were used in calculating criterion values. All involve
wild-caught adults taken from selenium contaminated streams and spawned for effects
determination. Exposure for all studies was therefore through the parents. Data are available for
rainbow trout (Oncorhynchus mykiss), cutthroat trout (1Oncorhynchus clarkii), Dolly Varden
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(,Salvelinus malma) and brown trout (Salmo trutta). The studies with Salvelinus are discussed
below; Oncorhynchus and Salmo were previously discussed in Section 3.1.2,
Salvelinus fontinalis (brook trout)
These data were not used directly in the criterion calculations. See Section 6.1.5 for
discussion of the available data.
Salvelinus malma (Dolly Varden)
Golder (2009) collected adult Dolly Varden from a reference site and two sites
downstream from the Kemess Mine in northern British Columbia, one with a high and one with a
moderate selenium exposure in the fall of 2008. Fertilized eggs were taken to the laboratory
where they were monitored for survival and deformities until 90% of the larvae reached swim-
up, approximately five months after fertilization. Alevin mortality was <1% in the treatments
collected from the exposed sites and not considered an effect. The prevalence of deformities
increased sharply after the selenium egg concentration exceeded 50 mg/kg dw (Appendix C).
The proportion of Dolly Varden larvae without any type of deformity (skeletal, craniofacial, and
finfold as well as edema), as a function of the log of the selenium concentration in the eggs using
TRAP, produced an EC io value of 56.22 mg Se/kg dw and an EC20 value of 60.12 mg Se/kg dw.
6.1.2.4	Salmonidae SMCV and GMCV Summary
As given in Section 3.1.2, the SMCV for cutthroat trout, Oncorhynchus clarkii, is 26.2
mg Se/kg dw in eggs derived from Rudolph et al. (2008), and Nautilus Environmental (2011),
(24.7, and 27.7 mg Se/kg dw respectively). The GMCV for the genus Oncorhynchus is 25.3 mg
Se/kg dw in eggs, derived from the 24.5 mg Se/kg dw EC10 from the combined Holm (2002) and
Holm et al. (2005) rainbow trout data, the above mean of the Rudolph et al. (2008) and Nautilus
Environmental (2011) Westslope cutthroat trout studies. The GMCV for the genus Salvelinus is
the EC10 value of 56.22 mg Se/kg dw for Dolly Varden (S. malma) from the Golder (2009) study.
6.1.2.5	Poeciliidae
Data are available for two species in this family. These studies are not represented in
Table 3.1 because they are live-bearing rather than egg-laying, but the relative tolerance of these
species is accounted for in derivation of the criterion.
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Gambusia holbrooki (eastern mosquitofish)
Staub et al. (2004) collected male and gravid female eastern mosquitofish from a
contaminated ash basin and a reference pond in July 1999. Male fish were used for measuring
standard metabolic rate and the reproductive endpoints. Brood size and percent viability of live
offspring at parturition were measured using the live-bearing females. Standard metabolic rates
of males, brood size of females, and offspring viability were not significantly different between
sites. Average concentrations of selenium in females were 11.85 and 0.61 mg/kg dw in the
contaminated ash basin and reference sites, respectively. The chronic value in whole body tissue
is >11.85 mg Se/kg dw whole-body (Appendix C). In a community of equally exposed fish taxa
(fish taxa having whole body tissue concentrations >11.85 mg Se/kg dw), the median egg-ovary
concentration among egg-laying fish would be expected to be 1.71 higher, or >20.26 mg Se/kg
dw.
Gambusia affinis (western mosquitofish)
Western mosquitofish were collected in June and July 2001 from sites in the grassland
water district in Merced County, California. Mosquitofish were collected from two sites that
were contaminated with selenium and from two reference sites in the same area with relatively
low selenium water concentrations (Saiki et al. 2004). Seventeen to 20 gravid females
(mosquitofish are live-bearers) from each site were held in the laboratory for two weeks to
quantify live and dead births and to make other measurements. Live and dead fry were visually
examined under low magnification with a binocular microscope for evidence of external
abnormalities (teratogenic symptoms such as spinal curvature, missing or deformed fins, eyes
and mouths and edema). The percentage of live births was high at both selenium-contaminated
sites (96.6 to 99.9%) and reference sites (98.8 to 99.2%). There were no obvious anomalies (e.g.,
deformities, edema) observed during the study. The concentration of selenium in four postpartum
females from the site with the highest selenium concentration ranged from 13.0 to 17.5 mg Se/kg
dw (geometric mean of the high and low is 15.1 mg Se/kg dw). The chronic value in whole body
tissue is >15.1 mg Se/kg dw (Appendix C). Similar to Staub et al. (2004), this value can be
converted to egg-ovary concentrations that would be expected in accompanying egg-laying fish,
by multiplying by the median fish egg-ovary to whole-body concentration ratio, 1.71. This yields
a >25.82 mg Se/kg dw equivalent egg-ovary concentration.
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Gambusia, which have been reported to be tolerant to selenium contamination, are often
one of the few remaining species at sites with high levels of selenium contamination (Cherry et
al. 1976; Lemly 1985a; Saiki et al. 2004; Young et al. 2010; Janz et al. 2010). The two studies
discussed above support this observation with a GMCV of >13.4 mg Se/kg dw in whole body
tissue, combining these "greater than" values as described in Section 6.1.1, It may be concluded
that this genus is not among the most sensitive to selenium.
6.1.2.6	Cyprinodontidae
Cyprinodon macularius (desert pupfish)
Besser et al. (2012), using a diet of oligochaete Lumbriculus that had fed on selenized
yeast, exposed desert pupfish to six levels of dietary and waterborne selenium. Five-week old
juveniles (Fo) were exposed for 85 days, during which time survival and growth were measured.
Upon reaching maturity at the end of this exposure period, the 60-day reproductive study was
begun, during which Fi eggs were collected, counted, and further tested for percent hatch,
survival, growth, and deformities. The authors observed no significant differences in pupfish
survival, growth, total egg production, hatch, or deformities among treatments. Although the
authors noted a potential interaction between the timing of egg production and treatment, a
comprehensive re-analysis of this data, described in Appendix C, indicated that the phenomenon
was neither statistically nor biologically significant. It is concluded that the egg concentration, 27
mg Se/kg (dw), for the test's highest treatment was not sufficiently high to define a
concentration-response curve. Although desert pupfish is thus not among the most sensitive
species, the slightly reduced survival observed at 27 mg Se/kg egg dw egg suggests that the ECio
may be close to that concentration, as also noted by the authors.
6.1.2.7	C entrarchi dae
Micropterus salmoides (largemouth bass)
A laboratory study was conducted by Carolina Power & Light (1997) in which adult
largemouth bass obtained from a commercial supplier were fed an artificial diet spiked with a
gradient of selenomethionine concentrations for several months. Approximately 100 eggs from
each spawn were monitored for mortality and deformities up to the larval swim-up stage. The
authors combined survival and deformities into a single metric (i.e., survival as normal
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offspring). The average concentration of selenium in the ovaries ranged from 3.1 mg/kg dw in
the control to 77.6 mg/kg dw in the high dietary treatment (53.1 mg/kg dw). A plot of the percent
survival of larval largemouth bass as a function of the selenium concentration in the parental
female ovary shows two groups of data; one at background survival with considerable variability
(mean 90.3%, standard deviation 10.9%) and one with <10% survival, with most of the data
being at 0% survival (see Microplerus summary in Appendix C, Figure 1). Due to inadequate
partial effects, a TRAP interpolation was used to estimate an ECio value. Based on a risk
management decision that the LOEC cannot be any higher than the lowest concentration with 0%
survival (32.9 mg/kg) and that any ECx should be below this, this establishes the higher
concentration point for the interpolation (an ECioo of 32.9 mg/kg) and requires that the highest 4
NOECs not be considered in setting the ECo. The lower concentration point for the interpolation
is therefore set here to 24.6, the next highest NOEC with greater than the average 90.3%
background survival. This results in an ECio of 26.3 mg/kg (and a steep slope of 16). Please see
Appendix C for more detailed information.
6.1.3 Reproductive Effects in Catfish (Ictaluridae)
Some important families of fish are not represented in the effects assessment, such as the
catfish family (Ictaluridae). In their compilation of egg-ovary versus whole-body ratios,
Osmundson et al. (2007) found comparatively high concentrations of selenium in egg-ovary
compared to whole body in black bullhead, Ameiurus melas, which are related to the Ictaluridae.
This raises a question about the potential risks of reproductive effects in this species and possibly
in related Ictaluridae. In addition to this concern about how much selenium such species may
accumulate in their eggs, U.S. Fish and Wildlife Service (2005) has suggested that offspring of
channel catfish (Ictaluruspunctatus) and related species might be affected at unusually low egg
concentrations. This is based on results of a study in which adult female catfish were injected
with seleno-L-methionine (Doroshov et al. 1992b). Effects were found in the offspring at egg
concentrations between 3.2 mg/kg (NOEC) and 6.3 mg/kg (LOEC), below levels observed in the
studies summarized in Section 3.1.2 and documented in Appendix C. These data were not
included in derivation of the criterion because the injection route of exposure is not an acceptable
experimental protocol for studies used in criterion derivation due to its difference from exposure
routes in the environment (water column and diet).
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In the absence of valid tests yielding an Ictaluridae ECio or chronic value, EPA evaluated
the potential vulnerability of the taxonomic group that includes catfish by examining
comparative fisheries observations of Ictaluridae and Centrarchidae sharing the same selenium-
contaminated water body. Crutchfield (2000) reports results of annual cove rotenone sampling
performed from 1982 to 1997 in Hyco Reservoir, North Carolina. The sampling was begun after
centrarchid populations in this reservoir had collapsed due to the release of ash pond selenium
from a coal-fired power plant. The plant began operating a dry fly ash handling system in
January 1990, thereby eliminating the aquatic discharge of selenium; the sampling continued
through the recovery period.
Crutchfield (2000) reports abundance data (kg/ha) for 20 fish taxa, including four
Ictaluridae and three Centrarchidae. These data were examined to determine the relationship
between the Ictaluridae and the selenium-affected Centrarchidae populations. The correlation
matrix between annual measured abundance of the seven taxa is shown below in Table 6.1.
Correlation with the reciprocal of measured average concentrations of selenium in invertebrates
is also shown. Because the reciprocal of the selenium concentration is used, a positive correlation
means that abundance decreases as selenium concentration increases. Conversely, a negative
correlation means abundance decreases as selenium concentration decreases.
Table 6.1. Correlation Matrix (Values of r) for Ictaluridae and Centrarchidae Abundance
and for Selenium Food Chain Contamination for the Hyco Reservoir.
(Data Reported by Crutchfield 2000).	

Ictaluridae
("enlrarchidae







1 .aiue-
I'omoxis


("humid
While
Hal
. \mcimns

111 CHI I ll
SPP
1 1 n\ erleh

catfish
catfish
hull head
spp
IJIueui
II hass
(crappie)
Se Cone
Channel catfish
1.00
-0.36
0.18
0.68
0.08
-0.33
-0.08
-0.44
White catfish
-0.36
1.00
0.02
-0.32
-0.31
-0.24
-0.15
-0.06
Black bullhead
0.18
0.02
1.00
0.40
0.32
-0.08
0.08
-0.03
Ameiurus spp.
0.68
-0.32
0.40
1.00
0.22
-0.24
-0.05
-0.31
Bluegill
0.08
-0.31
0.32
0.22
1.00
0.78
0.76
0.80
Largemouth bass
-0.33
-0.24
-0.08
-0.24
0.78
1.00
0.78
0.92
Pomoxis spp.
(crappie)
-0.08
-0.15
0.08
-0.05
0.76
0.78
1.00
0.69
1 Inverteb.
Se Cone.
-0.44
-0.06
-0.03
-0.31
0.80
0.92
0.69
1.00
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The centrarchid abundances are well correlated with each other and are closely related to
selenium concentrations in the food chain, with fish abundance decreasing as selenium
concentrations increase. Ictaluridae abundances, however, are unrelated either to the selenium-
sensitive centrarchid abundances or to the selenium concentrations in the food chain.
Figure 6.1 shows abundance as both mass and numbers of individuals of channel catfish
(CCF) and largemouth bass (LMB) observed by Crutchfield (2000) during the period 1982-1997.
Both species are long lived. Crutchfield (2000) notes that the decline of reproductive success and
abundance of Hyco's largemouth bass (and bluegill) was first documented in the mid-1970s.
Because this study was initiated after the largemouth bass recreational fishery had collapsed,
Figure 6.1 does not show the largemouth bass decline, only the period of its depression and
subsequent recovery.
Numbers of largemouth bass were very low at the beginning of the study period; their
numbers and mass do not begin to recover until invertebrate selenium drops below 30 mg Se/kg
dw. In the later portion of the study period, 1989-1997, largemouth bass numbers and mass
increase 100-fold. These observations are fully consistent with the premise that the earlier
observations of elevated selenium concentrations had been impairing reproduction of largemouth
bass.
In contrast, the ups and downs of channel catfish numbers, mass, and size seem to vary
randomly throughout the period of study. In 1984 catfish numbers reached their third highest
value while their average size was at its minimum: that is, there were many young individuals.
Simultaneously, largemouth bass was near its minimum for both numbers and mass. The next
year (1985) catfish numbers jumped to their maximum for the study period, and mass reached
near maximum. Such observations are easily explained if reproduction is taking place. But they
seem inexplicable under a premise that channel catfish reproduction was even more impaired
than largemouth bass reproduction, and its population merely a senescent non-reproducing
remnant of the pre-contamination population. Rather the observations indicate that if selenium
was having any effect on catfish reproduction, it was far less than on largemouth bass
reproduction and was no hindrance to rapid population increases.
Observations of selenium-contaminated Belews Lake accord with the above observations
of Hyco Reservoir. Young et al. (2010) indicate that as many as 29 resident fish species were
documented prior to contamination, but only common carp, catfish, and fathead minnows
113

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remained after contamination. The Doroshov et al. (1992b) injection study results suggesting that
channel catfish is sensitive at egg concentrations of 5 mg Se/kg dw, four-fold below the
largemouth bass Chronic Value, thus conflict with field observations. As demonstrated in the
Appendix C discussion of the Cleveland et al. (1993) toxicity tests with juvenile bluegill, the
exposure route by which selenium was accumulated can have a dramatic influence on the
potency of a given tissue concentration. That is, to accord with the Cleveland et al. (1993) data,
the whole-body EC50 would be expected to be at least four-fold higher when accumulated via
diet than when accumulated via water. For this reason, the criterion is derived only from tests
using the environmentally relevant exposure route of diet.
114

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400
300
200
100
0
CCF N/ha
LMB N/ha
/\ '
/ x-"

1982
1984 1986 1988 1990 1992 1994 1996
CCF kg/ha
LMB kg/ha
1982 1984 1986 1988 1990 1992 1994 1996
0.5
fi0.4
a>
N
In

<
0.3
CCF kg/N

1982
1984 1986
1988 1990
1992
1994 1996
txo
*3o
80
60

20
Tilapia All treatment
Reinitiation of bass
accidently processes
tournaments 1994.
introduced, operational,
CCF and LMB
1984 , early 1990
consumption ad-

visories lifted 1995.


Invertebrate Se Cone

i i i i i i i
1982
1984 1986 1988 1990
1992
1994 1996
Figure 6.1. Crutchfield (2000) Observations of Channel Catfish (CCF) and Largemouth
Bass (LMB) in Hyco Reservoir Beginning a Few Years after Populations of Largemouth
Bass had been Reduced by Se Contamination.
(A) number of individuals/ha, (B) mass/ha, (C) mass/ha divided by number/ha, yielding average
weight per individual, and (D) invertebrate Se concentration (mg Se/kg dw), and noting other
events relevant to management of the fishery.
115

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6.1.4	Reproductive Effects in Amphibians (Xenopus laevis)
Masse et al. (2015) has conducted the only maternal transfer study conducted with an
amphibian under controlled laboratory conditions. The African clawed frog {Xenopus laevis) was
fed a control diet (0.73 mg/kg Se dw) and three spiked diets containing selenium concentrations
of 10.92, 30.4 and 94.2 mg/kg dw. Trophic transfer to the frog's eggs was approximately 1:1
with measured selenium concentrations in the control and three spiked diets of 1.6, 10.82, 28.13,
and 81.66 mg/kg egg dw, respectively. Deformities were assessed in 200 tadpoles per female
(1800 - 2000 tadpoles per treatment group). ECio values determined by the authors for abnormal
spinal curvature, abnormal craniofacial structure and abnormal lens structure were 57.3, 38.4,
and 34.5 mg/kg Se egg dw, respectively. The ECio value for total deformities of 44.9 mg/kg Se
egg dw is in the upper-range of ECio values for fish (see Table 3.2). Although X laevis is a non-
native amphibian with a different reproductive strategy, their upper-range sensitivity suggests
amphibians are protected by the fish chronic criterion elements.
6.1.5	Reproductive Studies Not Used in the Numeric Criterion Derivation
Danio rerio (zebrafish)
Two studies (Penglase et al. 2014; and Thomas and Janz 2014) have shown the zebrafish,
Danio rerio (family Cyprinidae), to be sensitive to selenium. Penglase et al. (2014) assessed the
interaction of selenium with mercury through a maternal transfer study but did have two
treatments with selenium exposures resulting in 1.17 mg/kg egg dw (control) and 6.24 mg/kg egg
dw. The higher Se egg concentration had significantly reduced embryo survival and fecundity
relative to the control, however embryo survival in the controls was low at 54%. With only one
selenium treatment exposure, the data were not amenable to TRAP analysis. Thomas and Janz
(2014) conducted a maternal transfer study using adult zebrafish that were fed a control diet and
three levels of selenomethionine, 3.7, 9.6, and 26.6 mg/kg Se dw for 90 days before breeding the
exposed fish and collecting the fertilized embryos for assessment. TRAP analysis of larval
survival and larval deformities of 2-6 days post fertilization fish produced very low ECio values.
The lowest ECio was for deformities at 7.0 mg/kg egg dw. This value is markedly lower than any
of the ECio's in the current data set. The slope of the concentration-response curve for both
deformities and larval survival was very shallow, which was different than the selenium
responses for all other fish species for which data were available (see Figure E-6 in Appendix E).
Further, the control mortality in the experiment continuing over 160 days was high, over 40%.
116

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This zebrafish ECio for deformities contrasts with the absence of deformities in the
related species, fathead minnow, observed by GEI (2008) at concentrations up 40 mg/kg in adult
whole body (dw) as presented in Figure E-3 in Appendix E. The GEI (2008) fathead minnow
study was not directly used for criterion derivation because the offspring survival data for Sand
Creek appeared to be confounded by multiple stressors in this industrial waterway. However, its
deformity data appear unequivocal, indicating that the fathead minnow deformity endpoint is
relatively insensitive to selenium.
Since the zebrafish is a non-native cyprinid species, EPA assessed the information
available on zebrafish sensitivity to selenium compared to the sensitivity of native cyprinid
(minnow) species across the U.S. (Appendix E in the 2016 criterion document), including several
studies where native cyprinids were investigated in selenium-impacted waters (NAMC 2008).
Data from these studies suggest that native cyprinids are likely less sensitive to selenium than the
non-native zebrafish.
The anomalous nature of the concentration-response curve, with the very low value
coupled with field and other laboratory data suggesting that cyprinids are not particularly
sensitive to selenium was the basis for not including the zebrafish ECio in the data for deriving
the criterion. A detailed write up of this study and a summary of field and laboratory studies
indicating native cyprinids are not one of the more sensitive families are provided in Appendix
E.
Oncorhynchus clarkii (cutthroat trout)
Kennedy et al. (2000) reported no significant differences in mortality and deformity in
eggs, larvae, and fry from wild-caught cutthroat trout between a reference and an exposed site
(Fording River, British Columbia, Canada). The observations were made on eggs reared in well
water from spawning age females collected from the two locations (N = 17 and 20, respectively)
and fertilized by one male collected at each site. The mean selenium content in eggs from fish
collected from the reference site was 4.6 mg/kg dw and from fish collected from the Fording
River was 21.2 mg/kg dw. The chronic value for eggs is >21.2 mg Se/kg dw. These values were
not used in the criterion derivation because they represent high "greater than" values, as
discussed above, and provide no additional important quantitative data for the analyses.
Hardy (2005) fed cutthroat trout experimental diets containing a range of
selenomethionine (0-10 mg/kg dw) for 124 weeks. No significant growth or survival effects were
117

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observed in the adult fish over the 124 weeks. The whole body concentration reached 12.5 mg/kg
dw selenium after 44 weeks. Embryo-larval observations (percent hatch and percent deformed)
were not related to whole body selenium concentrations in the spawning females (9.37 mg/kg
dw) fed the selenium-laden diet for 124 weeks. The concentration of selenium in eggs from these
females was 16.04 mg/kg dw. For this study the chronic value, an unbounded NOEC, is thus
>16.04 mg Se/kg dw in eggs. This value was not used in the criterion derivation.
Salvelinus fontinalis (brook trout)
Holm et al. (2005) collected spawning brook trout from streams with elevated selenium
contaminated by coal mining activity and from reference streams in 2000, 2001 and 2002.
Similar to procedures described by these authors for rainbow trout, above, fertilized eggs were
monitored in the laboratory for percent fertilization, deformities (craniofacial, finfold, and spinal
malformations), edema, and mortality. Embryos from the contaminated stream had on average a
higher frequency of craniofacial deformities than fry from the reference stream (7.9% for the
contaminated stream compared to 2.1% in the reference stream). Although this increased rate of
craniofacial deformities was calculated to be statistically significant when compared across sites,
the Abbott-adjusted effect is only 6% and is thus below the 10% effect represented by an ECio.
But more important, when comparing across adult females (the more reliable analysis for
selenium reproductive toxicity studies of this type, and the one used to obtain the related rainbow
trout ECio for these authors' studies), there is no apparent relationship between brook trout
craniofacial deformities and exposure across a broad range of concentrations, as illustrated in
Appendix C. An environmentally conservative estimate of the NOEC might be considered to be
the average concentration of selenium in eggs from the high exposure site (Luscar Creek), >7.78
mg Se/kg ww or >20.5 mg Se/kg dw using the 61.2% moisture content for rainbow trout eggs
cited above. However, the effect threshold appears to be substantially higher based on the
absence of any consistent concentration-response relationship up to the maximum observed egg
concentration of 18.9 mg Se/kg ww or 48.7 mg Se/kg dw, as shown in the Appendix C graphs.
Given the point estimate ECio available for the related species, Salvelinus malma (Dolly Varden,
Section 6.1.2.3), the "greater than" chronic value for brook trout is not used to obtain the
GMCV, in accordance with the principles listed in Section 6.1.1,
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Lepomis macrochirus (bluegill)
Applicable chronic reproductive data for bluegill can be grouped by exposure type: field
and laboratory. In some field studies, chronic value estimates were "less than" fairly high
selenium concentrations (Bryson et al. 1984, 1985a; Gillespie and Baumann 1986). This low
resolution is due to the observed effect occurring at a single observed high exposure
concentration relative to a reference condition. In the Bryson et al. (1984, 1985a) and Gillespie
and Baumann (1986) studies, the artificially crossed progeny of females collected from a
selenium contaminated reservoir (Hyco Reservoir, Person County, NC) did not survive to swim-
up stage, irrespective of the origin of milt used for fertilization. Measured waterborne selenium
concentrations prior to the experiments ranged from 35 to 80 |ig/L. The ovary tissue selenium
concentration associated with this high occurrence of mortality of hatched larvae was <30 mg/kg
dw tissue, as reported by Bryson et al. (1985a), and <46.30 mg/kg dw tissue, as reported by
Gillespie and Baumann (1986). In the case of the latter, nearly all swim-up larvae from the Hyco
Reservoir females were edematous, none of which survived to swim-up.
Bryson et al. (1985b) examined percent hatch and percent swim-up larvae from spawns
using bluegills collected from Hyco Reservoir and a control site. There were no differences in
the Hyco measurements relative to the control. The concentration of selenium in the liver of the
parental Hyco bluegill was 18.6 mg/kg dw. The chronic values for this embryo-larval
development test was >18.6 mg Se/kg dw liver. The high "less than" and low "greater than"
chronic values obtained from Bryson et al. (1984, 1985a, b) and Gillespie and Baumann (1986)
were not used in the SMCV calculation because these values are consistent with and yet provide
no numeric basis for modifying the SMCV obtained from the ECioS.
6.1.6 Salmo GMCV: EPA Re-analysis of a Key Study Used in Criterion Derivation
Previously, in the draft selenium criterion document submitted for external peer review in
May 2014, the lowest GMCV in the reproductive effects dataset was for Salmo (15.91 mg/kg
dw) based on larval deformities. Subsequently, in 2015, EPA conducted a careful re-analysis in
response to stakeholder comments to confirm the validity of the approach used in 2014, resulting
in the calculation of an ECio at 18.09 mg/kg dw based on larval mortality from hatch through
swim up, prior to a lab overflow accident during the post swim up feeding portion of the test.
The dataset was constrained to hatch through swim up information due to uncertainty introduced
by the loss of larvae from an overflow event caused by clogged drains during the post swim-up
119

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portion of the test (Formation 2011). The hatch through swim up deformity endpoint was not
considered because of the preferential selection of visibly non-deformed fish for the post swim
up portion of the test (Response letter to EPA, J.R. Simplot Company 2014). This is important
because the primary endpoint of interest during the post-swim up phase is deformity rate.
Random selection of living fish would have been more appropriate since visibly healthy fish may
be less likely to express deformities in this later stage of the test.
Following the release of the 2015 draft selenium criterion document, the larval survival
from hatch through swim up dataset was reanalyzed, and it was determined that the TRAP model
resulting in an ECio of 18.09 mg/kg was not appropriate, because the ECio was lower than one of
the treatment levels within the background no-effect range. In order to calculate an ECio that
would not fall below the highest background concentration, a weighted least squares nonlinear
regression was calculated in TRAP, resulting in an ECio of 21.0 mg/kg. Additional details
describing this weighted nonlinear regression approach are described in Appendix C.
6.1.7	Impact of Number of Tested Species on Criterion Derivation
Many of the species used for testing the toxicity of selenium are those observed to be
affected at contaminated sites or otherwise suspected to be particularly sensitive. Six of the eight
minimum data requirements were met, and the other two (for planktonic and benthic crustaceans)
were waived (see Section 2.6). Of the N=15 genera used for the calculation of the criterion, ten
are fish, which are more sensitive than invertebrates, based on the available data. Of the ten fish
genera, five are either salmonids or centrarchids. Had a broader array of expected insensitive
taxa been included, the four most sensitive genera would not likely change, but N would
increase. The criterion calculation for selenium is relatively insensitive to the effect of increasing
the value of N by adding more tests with different genera than those already represented. Setting
N=20 (leaving the four most sensitive the same) would only raise the egg-ovary criterion element
from 15.1 mg Se/kg to 16.0 mg Se/kg. This insensitivity occurs because the four lowest GMCVs
are closely spaced, such that the calculated egg-ovary criterion element is never distant from the
lowest GMCV.
6.1.8	Comparisons between Concentrations in Different Tissues
Researchers often report concentrations of selenium in fish eggs or ovaries (e.g.,
Formation Environmental 2011, 2012; Holm et al. 2005; Osmundson et al. 2007). Osmundson et
120

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al. (2007) found reduced levels of selenium in ovaries after spawning, presumably due to the loss
of selenium through spawning and release of eggs with relatively high concentrations of
selenium. Of the 14 chronic values determined from the maternal transfer reproductive studies,
12 values represent selenium measured in eggs. Two values represent selenium measured in the
ovaries: Hermanutz et al. (1992, 1996) and Carolina Power & Light (1997). Hermanutz et al.
(1992, 1996) sampled adult female bluegill just prior to spawning and at the end of the test (post
spawning) and found no decreases in the concentration of selenium in the post-spawned fish. In
the Carolina Power & Light (1997) study, selenium in ovaries of largemouth bass was measured
from fish sampled just after spawning. No comparison to prespawning fish or selenium in eggs
can be made for the largemouth bass study, however, the ECio of 26.3 mg Se/kg ovary dw was
mid-range in the SSD indicating this test was not overly conservative due to lower selenium
measurements in post spawning ovaries. Based on the observations stated above, egg selenium
and ovary selenium were considered equal for the toxicity data set. Any potential error resulting
from this assumption would be conservative since the effect of spawning only lowers the
selenium concentration in the ovary. EPA recognizes selenium ovary concentrations may vary in
field collected samples due to fish reproductive cycles and will address such concerns in the
implementation information.
6.1.9 Studies of Non-Reproductive Effects
Non-reproductive effect studies do not involve effects on the offspring of exposed female
adults, and their results are not expressed as selenium concentrations in egg or ovary tissue.
Because selenium concentrations in whole body and muscle are generally lower than in egg and
ovary, with observed egg-ovary to whole-body ratios ranging from 1.3 to 7.4, and egg-ovary to
muscle ratios ranging from 1.0 to 5.8, whole-body, muscle, and egg-ovary effect concentrations
can only be compared after accounting for the inherent differences in the selenium
concentrations in these different media. Non-reproductive effects were determined to provide a
less reliable basis for a criterion, in part because comparatively few of such studies provided
sigmoidal concentration-response curves. Non-reproductive SMCVs and GMCVs are shown in
Table 6.2 below and summaries of the acceptable non-reproductive studies are included in
Appendix D.
121

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Table 6.2. Freshwater Chronic Values from Acceptable Tests - Non-Reproductive Endpoints.
(Parental Females Not Exposed).	


l'l\poMirc mule

Toxicologic;!!
Chronic Y;iluc
SMCY
CM( V
Species
Reference
iiiid (lui'iilioii
Selenium Itiriii
cndpoinl
niii/kii (In ''
m;*/kg (Im
nig/kg (In
Acipenser
transmontanus
white sturgeon
Tashjian et al.
dietary (lab)
seleno-L-methionine in
artificial diet
ECiojuvenile growth
15.08 WB
27.76 M
ECio
15.1 WB
27.8 M
15.1 WB
2006
8 weeks
seleno-L-methionine in
artificial diet
EC2o juvenile growth
17.82 WB
32.53 M
ec20
17.8 WB
32.5 M
27.8 M




NOEC
10.1 M


Pogonichthys

dietary (lab)
9 months

LOEC
15.1 M
10.1 M
10.1 M
macrolepidotus
Sacramento splittail
Teh et al. 2004
selenized-yeast
MATC juvenile
deformities (juvenile
exposure only)
12.34 M
15.1 M
12.3 M
15.1 M
12.3 M
Pimephales promelas
fathead minnow
Bennett etal. 1986
dietary (lab)
9 to 19 days
algae exposed to selenite
then fed to rotifers which
were fed to fish
Chronic value for
larval growth
51.40 WB
51.40 WB
69.83 M
51.40 WB
69.83 M
Pimephales promelas
fathead minnow
Dobbsetal. 1996
dietary and
waterborne
(lab)
8 days
algae exposed to selenate
in water then fed to
rotifers which were fed
to fish
LOEC for larval fish
dry weight after 8 d
<73 WBb



water: selenate; diet:




Xyrauchen texanus
razorback sucker
Beyers and
Sodegren 2001a
dietary and
waterborne (lab)
28 days
algae exposed to selenate
in water then fed to
rotifers which were fed
to fish
NOEC for survival
and growth
>12.9 WBb
see text
see text
Xyrauchen texanus
razorback sucker
Beyers and
Sodegren 2001b
dietary and
waterborne (lab)
28 days
water: site waters; diet:
algae exposed to site
water then fed to rotifers
which were fed to fish
NOEC for survival
and growth
>42 WBb





water: selenate;




Catostomus latipinnis
flannelmouth sucker
Beyers and
Sodegren 2001a
dietary and
waterborne (lab)
28 days
diet: algae exposed to
selenate in water then
fed to rotifers which
were fed to fish
NOEC for survival
and growth
>10.2 WB
>10.2 WB
>10.2 WB
Oncorhynchus
tshawytscha
Hamilton et al.
1990
dietary (lab)
60 days
mosquitofish spiked with
seleno -DL-methionine
ECio for juvenile
growth
7.355 WB
ECio
9.052 WB
ECio
9.052 WB
122

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Species
Keleivnce
l'l\poMirc mule
iiiid (lui'iilioii
Selenium I n nil
Toxicologic;!!
cndpoinl
C hronic Y;iluc
niii/kii (In ''
SMCY
nii*/kK cl
CM( V
m/k*i (l\\
chinook salmon



EC20 for juvenile
growth
10.47 WB
ec20
12.83 WB

mosquitofish spiked with
SLD diet
ECio for juvenile
growth
11.14 WB
EC20 for juvenile
growth
15.73 WB
Oncorhynchus mykiss
rainbow trout
Hilton and Hodson
1983;
Hicks et al. 1984
dietary (lab)
16 weeks
sodium selenite in food
preparation
juvenile growth
NOEC
21 Liver
NOAEC
28.98 L
LOAEC
84.68 L
MATC
49.52 L
LOEC
71.7 Liver
MATC
38.80 Liver
Oncorhynchus mykiss
rainbow trout
Hilton etal. 1980
dietary (lab)
20 weeks
sodium selenite in food
preparation
juvenile survival and
growth
NOEC
40 Liver
LOEC
100 Liver
MATC
63.25 Liver
Morone saxitilis
striped bass
Coughlan and
Velte 1989
dietary (lab)
80 days
Se-laden shiners from
Belews Lake, NC
LOEC for survival of
yearling bass
<16.2 Mc
<16.2 M
<16.2 M
Lepomis macrochirus
bluegill
Lemly 1993a
dietary and
waterborne (lab)
180 days
20 to 4C
diet: seleno-L-
methionine
water: 1:1
selenate: selenite
LOEC for juvenile
mortality at 4C
<7.91 WB
4C
ECi o-NO AEC
8.15 WB
4C
ec20-loaec
8.80 WB
9C ECio
14.0 WB
9C EC20
14.6 WB
4C & 9C
9.33 WB
Threshold prior to
"winter stress"
5.85 WB
dietary and
waterborne (lab)
180 days 20C
diet: seleno-L-
methionine
water: 1:1
selenate: selenite
NOEC for juvenile
mortality at 20C
>6.0 WB
Lepomis macrochirus
bluegill
Mclntyre et al.
2008
dietary and
waterborne (lab)
182 days
20 to 4C (ESI)
diet: Lumbriculus fed
selenized-yeast
water: 1:1
selenate: selenite
ECiojuv. survival
ESI
9.27 WB
EC20juv. survival
ESI
9.78 WB
dietary and
waterborne (lab)
182 days
20 to 9C (ES3)
diet: Lumbriculus fed
selenized-yeast
water: 1:1
selenate: selenite
ECiojuv. survival
ES3
14.00 WB
EC2ojuv. survival
ES3
14.64 WB
dietary and
waterborne (lab)
182 days
20 to 4C (ES2)
diet: seleno-L-
methionine
water: 1:1
selenate: selenite
NOEC juv. surv. ES2
>9.992 WB
123

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Species
Reference
l'l\poMirc mule
iiiid (lui'iilioii
Selenium Itiriii
Toxicologic;!!
cndpoinl
Chronic Y;iluc
niii/kii (In ''
SMCV
m;*/kg (Im
c;mcy
nig/kg (In
Lepomis macrochirus
bluegill
Bryson et al.
1985b
dietary (lab)
60 days
seleno-DL-cysteine
NOEC for juvenile
growth
>3.74 WBb


Lepomis macrochirus
bluegill
Cleveland et al.
1993
dietary (lab)
90 days
seleno-L-methionine
NOEC for juvenile
survival
>13.4 WBb
All chronic values reported in this table are based on the measured concentration of selenium in whole body (WB), muscle (M)
or liver (L) tissues.
Chronic value not used in SMCV calculation (see text).
Tissue value converted from ww to dw. See Appendix C for conversion.
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6.1.9.1 Comparison of Fish Chronic Reproductive Effects and Chronic Non-Reproductive
Effects
A chronic criterion element concentration of 15.1 mg/kg dw in the egg/ovary addresses
the toxic effect identified by the Chapman et al. (2009, 2010) expert workshop to be of greatest
concern, reproductive effects, and is expected to be protective of non-reproductive endpoints
such as juvenile survival and growth.
If the information in the reproductive-effect GMCV Table 3.2 (expressed as whole-body)
were combined with the information in the nonreproductive-effect Table 6.2, and the lower of
the reproductive or nonreproductive GMCVs for each tax on were used to construct a combined
distribution of whole-body chronic values, the resulting criterion element (corresponding to
N=18, accounting for three additional fish genera only having nonreproductive-effect GMCVs),
the FCV would be calculated to be 9.1 mg Se/kg WB dw, similar to the 8.5 mg Se/kg WB dw
FCV for reproductive effects expressed as whole-body Figure 6.2.
256
128

T3
0
CO
1
_0J
o
64 L
32
16
	Fish (TL3) WB Repro GMCV
~	Fish (TL3) WB Nonrepro GMCV
~ Inverteb (TL2) GMCV
Fish (TL3)WB FCV
t-JL-I
6
Rank
10
12
Figure 6.2. Distribution of Fish Reproductive Effect GMCVs from Figure 3.2 and
Distribution of Fish Nonreproductive Effect GMCVs and Invertebrate GMCVs.
For establishing a reliable criterion, the sufficiency of and consistency among the data
underlying the reproductive-endpoint GMCVs favor their use over any non-reproductive
endpoint data (see Section 3.1.1 and Appendix C). Most of the reproductive studies involved
125

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examining the offspring of wild-caught females, exposed under real-world conditions. Most had
concentration-response curves that supported ECio estimates.
In contrast, the non-reproductive endpoint studies provide fewer data for supporting a
criterion, and fewer of these studies yielded the type of concentration-response data that could
support EC 10 estimates. Furthermore, the non-reproductive data are not as consistent, as noted by
Janz et al. (2010). The reproductive effect data also show more clear-cut concentration-response
relationships than the non-reproductive effect data (11 of the reproductive chronic values are
specific ECs, compared to only five of the non-reproductive chronic values), are more readily
reproducible, and are better corroborated by field observations. Reproductive effects represent
the endpoint of greatest concern (Chapman et al. 2009, 2010); all non-reproductive GMCVs are
protected by a criterion derived from the reproductive GMCVs. The reproductive endpoint data,
expressed relative to selenium concentrations in fish eggs and ovaries, thus provide a more
reliable and protective basis for the criterion. Because the data set used to derive the criterion is
comprised primarily of the aquatic species considered most sensitive to selenium (salmonids and
centrarchids) and because the criterion is designed to protect 95% of the genera, the egg-ovary
criterion element concentration of 15.1 mg/kg dw ovary/egg should be protective of aquatic
populations of fish and invertebrates.
6.1.10 Special conditions for consideration of primacy of water column criterion elements over
fish tissue criterion elements
The chronic selenium criterion is derived to be protective of the entire aquatic
community, including fish, amphibians, and invertebrates. Fish are the most sensitive taxa to
selenium effects. Selenium water quality criterion elements based on fish tissue (egg-ovary,
whole body, and/or muscle) sample data supersede the criterion elements based on water column
selenium data, when measured in the same approximate time frame (approximately one year) and
site. This is due to the fact, noted above, that fish tissue concentrations provide the most robust
and direct information on potential selenium effects in fish. However, because selenium
concentrations in fish tissue are a result of selenium bioaccumulation via dietary exposure, there
are two specific circumstances where the fish tissue concentrations do not fully represent
potential effects on fish and the aquatic ecosystem: 1) In "fishless" waters, and 2) new selenium
inputs. Fishless waters are defined as waters with insufficient instream habitat and/or flow to
support a population of any fish species on a continuing basis, or waters that once supported
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populations of one or more fish species but no longer support fish (i.e., extirpation) due to
temporary or permanent changes in water quality (e.g., due to selenium pollution), flow or
instream habitat. Because of the inability to collect sufficient fish tissue to measure selenium
concentrations in fish within such waters, water column concentrations will best represent
selenium levels required to protect aquatic communities and downstream waters in such areas.
New inputs are defined as new activities resulting in selenium being released into a lentic or lotic
waterbody. New inputs will likely result in increased selenium in the food web, resulting in
increased bioaccumulation of selenium in fish over a period of time until the new selenium
release achieves a quasi-"steady state" balance within the food web. EPA estimates that
concentrations of selenium fish tissue will not represent a "steady state" for several months in
lotic systems, and longer time periods (e.g., 2 to 3 years) in lentic systems, dependent upon the
hydrodynamics of a given system; the location of the selenium input related to the shape and
internal circulation of the waterbody, particularly in reservoirs with multiple riverine inputs; and
the particular food web. Estimates of time to achieve steady state under new selenium input
situations are expected to be site dependent, so local information should be used to better refine
these estimates for a particular waterbody. Thus, EPA recommends that fish tissue concentration
not supersede water column concentration until these periods of time have passed in lotic and
lentic systems, respectively, or until steady state concentrations can be determined.
6.2 Water
6.2.1 Validation of Translation Equation for Developing Water Column Concentrations
EPA evaluated the efficacy of the equation used to translate the egg-ovary criterion
element to a water column concentration. EPA's translation equation is given as:
C
q 		egg-ovary	
"*' ~ TTF"" xEFxCF	(Equation 18)
Because selenium levels in fish are relatively stable over a long time period if the ecosystem is at
steady state with respect to selenium concentration, single measurements of selenium in fish
tissue are likely to be less variable and a better representation of selenium loads to the aquatic
system than single measurements of selenium in the water column. Thus, EPA used a validation
approach based on fish tissue measurements rather than single water measurements.
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Rearranging Equation 18 to solve for egg-ovary concentration yields:
c,s-0v = Cr x TTFx EF x CF	(Equation 22)
EPA used Equation 22 to calculate the predicted concentration of selenium in the eggs and
ovaries of fish from all spatially and temporally relevant measurements in the water column.
EPA then compared those predicted values to the measured concentration in the fish.
EPA searched its collection of selenium measurements in fish tissue taken from aquatic
sites with a previously calculated EF value. Identified tissue measurements from other than eggs
or ovaries were converted to equivalent egg-ovary concentrations using species-specific
conversion factors as described previously. For each tissue measurement, EPA searched its
collection of selenium measurements again for water column measurements that were taken from
the same aquatic site and within one year of the tissue measurement. If more than one water
column measurement was matched to a tissue measurement, the median water column
measurement was used. For each matched pair of tissue and water measurements, appropriate
species-specific TTF and CF values were identified as described previously, and the EF value
from the site samples were taken. EPA then used Equation 22 to calculate the predicted egg-
ovary concentration from the observed water column concentration. Finally, EPA compared the
predicted egg-ovary concentrations with the observed egg-ovary concentrations.
EPA identified 317 tissue measurements associated with one or more water column
measurements. A predicted egg-ovary concentration was calculated for each water column
concentration as described above. Figure 6.3 shows all 317 predicted egg-ovary concentrations
plotted against the measured egg-ovary concentrations. Because both the predicted and observed
selenium concentrations exhibited substantial heteroscedasticity (the variability of one variable is
unequal across the range of values of a second variable that predicts it), they are plotted and
analyzed on a log scale. The predicted and measured concentrations are highly correlated
(r=0.82, t(3i5)=25.30, P<0.001). Data used to generate Figure 6.3 can be found in Appendix I.
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1000.0
Observed
egg-ovary
concentration
(mg/kg dw)
100.0
10.0
1.0
0.
Ega-ovarv criterion element (mg/kg dw)
4
Egg-ovary
criterion
element
(mg/kg dw)
0.1	1.0	10.0 100.0 1000.0
Predicted egg-ovary concentration (mg/kg dw)
Figure 6.3. Scatter Plot of Predicted Versus Measured Concentrations of Selenium in Fish.
Solid line shows unity y = x line; dashed lines show the egg-ovary criterion element value.
Although there is a strong correlation between predicted and observed egg-ovary
concentration values, Figure 6.3 shows more data points above the y = x line at low selenium
concentrations. This result suggests the model underestimates bioaccumulation at low selenium
concentrations. Such behavior is likely the result of the inherent model assumption of constant
bioaccumulation rates regardless of selenium concentration, whereas selenium bioaccumulation
has been shown to be inversely related to water concentration (see Sections 3.2.2 and 3.2.4 for
further discussion). Within the range of concentrations near the egg-ovary criterion element
value, however, the relationship between predicted and observed selenium concentrations are
evenly dispersed around the y = x line. Thus the model is unlikely to result in biased estimates
near egg-ovary concentrations that may require regulatory action.
Dispersion around the unity line is likely attributable to several sources of uncertainty
including small sample sizes, temporal or spatial variability in selenium exposure, and local
variability in aquatic food webs. EPA limited this analysis to only those aquatic sites with at least
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two particulate measurements available to calculate an EF value and with at least one of them
from algae or detritus. The requirement of at least two particulate measurements was made
because a single measurement was considered insufficient. The requirement that at least one of
the measurements be for algae and/or detritus was made because selenium within these
particulate types was more highly correlated to water (Section 3.2.2.3), Nevertheless, only one or
two measurements of algae and/or detritus were available for 62 of the 96 aquatic sites evaluated.
Although the minimum data requirements described above reduce uncertainty when applying
Equation 22 to available data, EPA believes that two particulate measurements are only
marginally sufficient. Another potential source of uncertainty is the frequent absence of site-
specific information about the types and proportions of organisms ingested by fish. In most
cases, EPA estimated the type and proportion of prey organisms using general knowledge of the
fish species and aquatic system location. Notwithstanding the limitations in available data, the
EPA concludes from this analysis that Equation 18 provides a reasonable translation of the egg-
ovary criterion element to a site-specific water concentration.
6.2.2 Sulfate-Selenium Interactions
Several investigators (Brix et al. 2001; Ogle and Knight 1996; Williams et al. 1994) have
previously evaluated the role of sulfate on the bioavailability and toxicity of selenium in
freshwater organisms. A report from DeForest et al. (2014) notes that a sulfate-dependent
selenium criteria would apply only to selenate-dominated, well-oxygenated streams, a subset of
freshwater systems in the U.S. The DeForest publication discussed experiments to assess
influence of sulfate on selenate uptake on one species of macrophyte (Lemna minor) and one
algal species (Pseudokirchnella subcapitata), a limited data set of primary producers. The
authors note that, "It does need to be emphasized here, however, the analysis currently does not
include Se data for periphyton and benthic diatoms, as these data are not available." The authors
also note that, "due to methodological challenges and high costs, it is difficult to
comprehensively evaluate the influence of sulphate on bioconcentration and transfer up the food
chain."
Including any type of sulfate relationship in the national criterion derivation would
necessitate having sulfate measurements to accompany all observed selenium water
concentrations included in the derivation database. That is, the absence of an accompanying
sulfate observation would necessitate excluding the water observation. The resulting reduction in
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the number of sites included in the database would reduce the confidence in its ability to
represent the nation's waters. For the above reasons, EPA has not included a sulfate relationship
in the 2016 selenium criterion.
6.3 Uncertainty
This section examines several areas where EPA addressed uncertainty in the development
of the selenium water quality criterion. This section represents a qualitative treatment of specific
parts of the derivation process for the selenium freshwater chronic criterion where EPA has
identified the potential for uncertainty, and also describes the approaches that the Agency used to
reduce uncertainty.
EPA developed a tissue-based water quality criterion designed to be protective of aquatic
life from the chronic effects of selenium. In general, EPA followed the procedure detailed in the
document, Guidelines for Deriving Numerical National Water Quality Criteria for the Protection
of Aquatic Organisms and Their Uses (hereafter referred to as the "Guidelines") (Stephan et al.
1985). The Guidelines sets forth a methodology for deriving ambient water quality criteria for
the protection of aquatic life that includes a rigorous list of data quality requirements. Because
selenium is a bioaccumulative chemical with maternal diet and transfer as the primary route of
exposure for chronic toxicity, EPA included additional data quality requirements such as the
requirement of a dietary exposure. See Section 2.7.5 Analysis Plan for Derivation of the Chronic
Fish Tissue-Based Criterion Elements for how chronic effect levels were determined for
selenium. The Guidelines provide several recommended approaches that reduce uncertainty in
the derivation of criterion. It provides a strict set of guidelines for the acceptance of data to be
used in criteria derivation. It provides a minimum set of data requirements (MDRs) that define an
assemblage of aquatic organisms that can be used in a genus sensitivity distribution to derive a
criterion that is protective of 95% of aquatic species. The requirements in the Guidelines reduce
the uncertainty in the ability of a criterion to be protective of aquatic life.
6.3.1 Tissue Criterion Element
The tissue criterion element is based on reproductive effects caused by selenium and is
expressed in three different tissue types, egg/ovary, muscle and whole body. Non-reproductive
effects were also determined but not used in the derivation of the criterion because of less
certainty in the endpoints and effect levels (Section 6.1.9 and Table 6.2). A comparison of fish
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reproductive and non-reproductive effects and the conclusion that the reproductive criterion is
protective of the non-reproductive effects is given in Section 6.1.9.1,
The dataset used to derive the tissue-based criterion consists primarily of fish species: 12
fish species representing 10 genera and 7 families. Although this might be viewed as a small
number of the nearly 800 native freshwater fish species (36 families) of fish in the United States,
it is a large number of species relative to chronic criteria derivations for other pollutants.
Furthermore, the fish species that have been the focus of some of the research have been the
species observed to be those first affected (most sensitive) at selenium-contaminated sites such
as Belews Lake and Hyco Reservoir, i.e., bluegill sunfish and largemouth bass. The data set
contains three acceptable chronic toxicity studies with bluegill sunfish, the second most sensitive
genus in the dataset and one with largemouth bass, the fifth most sensitive genus. The three
replicate chronic values for bluegill are 14.7, 22.6, and 26.3 mg Se/kg egg-ovary dw. Of these
three values, 14.7 mg Se/kg ovary dw is likely the least certain because the study (Hermanutz et
al. 1992, 1996) was not designed to minimize uncertainty in characterizing the tissue
concentrations associated with its observed levels of effect.
Three genera representing five species from Salmonidae, a family considered to be
generally sensitive to contaminants, are in the data set. Two salmonid genera, Salmo and
Oncorhynchus are the third and fourth most sensitive taxa in the data set. Salmo is represented by
a single study, but because the study included a large number of individuals across a broad
spectrum of exposure, the uncertainty associated with its 21 mg Se/kg egg dw might not be
viewed as particularly large. The chronic values for the three studies with Oncorhynchus are
confined to the narrow range 24.5 - 27.7 mg Se/kg egg dw, and on that basis may be considered
to have low uncertainty. Although the numbers of species and families of fish in the data set are
a fraction of what are native to the United States, the fish species contained in the data set are
known to be those most sensitive to selenium based on field observations or known to be
sensitive in general to contaminants. With the lowest six chronic values falling in the relatively
narrow range 15.6-27 mg Se/kg egg-ovary dw, the selenium tissue criterion element should
probably be considered to have the smallest amount of uncertainty among the existing aquatic
life criteria.
As stated in the previous paragraph, the data set primarily consists of fish species and
contains only three invertebrate species. The cases in the field where adverse effects have been
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observed to fish and water birds (e.g., Belews Lake, Hyco Reservoir, Kesterson Reservoir) have
not documented any adverse effects on macroinvertebrates either on a species or community
level (Janz et al. 2010). The effect levels determined for the three invertebrate species contained
in the data set are consistent with the field observations that macroinvertebrates are in general
less sensitive to selenium than fish species. EPA recognizes that there may be more sensitive
oviparous taxa (fish and amphibians), as well as macroinvertebrate taxa than those in the current
data set and supports the testing of different species.
6.3.1.1	Reproductive Endpoints
Reproductive endpoints were determined from studies in which adults were exposed to
selenium either in the laboratory or field. Effects were measured in the offspring which received
selenium exposure via maternal transfer. Larval mortality and teratogenic deformities such as
skeletal, craniofacial, and fin deformities, and various forms of edema that result in mortality are
the most sensitive indicators of selenium toxicity in fish larvae. Recent research suggests the
mode of action of selenium-induced toxicity in fish larvae is due to oxidative stress and appears
to be related to glutathione homeostasis (See Section 2.3 for more detail on this subject.).
Linking the mode of action directly to the assessment endpoint used in the derivation of the
tissue-based criterion provides a consistent concentration-response relationship among the
studies used in the data set. Using the most sensitive assessment endpoint (based on the state of
the science) reduces uncertainty in the ability of the criterion to protect aquatic life.
6.3.1.2	Egg Ovary Chronic Values
Chronic Values (CV) were based on the most direct representation of exposure to the
effect in the offspring, that is, the concentration of selenium in the egg/ovary. One way to assess
the precision of the chronic values used in the derivation of the criterion is to look at the
reproducibility of tests used to calculate the CV for a taxon. This precision assessment can be
done with two tests conducted with cutthroat trout and three tests conducted with bluegill
sunfish. The two cutthroat trout studies (Rudolph et al. 2008 and Nautilus Environmental 2011)
had very similar ECio values (24.7 and 27.7 mg Se/kg egg dw, respectively) for the same
endpoint (larval survival) using fish collected at the same site. Two of the three bluegill tests
(Coyle et al. 1993 and Doroshov et al. 1992a) also had very similar ECio values (26.3 and 22.6
mg Se/kg egg dw, respectively) for larval endpoints determined in laboratory exposures. An
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ECio of 14.7 mg Se/kg ovary dw was determined in the third bluegill test, a mesocosm exposure
study reported by Hermanutz et al. (1992, 1996). Although the mesocosm study had a lower
ECio value when compared with Coyle et al. (1993) and Doroshov et al. (1992a), it was within a
factor of 1.8 and 1.5, respectively. Delos (2001) found such differences to be typical when
equivalent toxicity tests of the same species are compared. The relatively low variability between
chronic toxicity tests conducted with the same species indicates precision in the CV estimates,
which reduces the uncertainty the tissue-based criterion.
Most of the CVs were determined using an ECio value and a few were estimated using
the NOEC. ECio values were considered more appropriate than EC20, because selenium is a
tissue-based criterion due to its nature of exposure and effects for this bioaccumulative chemical.
See Section 2.7.1 for a discussion of why EC10S were favored over EC20S. The use of EC10S and
NOECs increases the certainty that the criterion will be protective of aquatic life.
6.3.1.3	Whole Body and Muscle Chronic Values
Effect levels (ECio or NOECs) were determined directly for whole body or muscle
tissues when the selenium concentrations for these tissues were measured and reported in the
tests. Effect levels were calculated directly using muscle tissue for five of the chronic toxicity
tests: northern pike, cutthroat trout (Rudolph), bluegill (Doroshov and Hermanutz) and white
sturgeon, while effect levels for three tests were calculated directly using whole body selenium
concentrations: bluegill (Coyle and Hermanutz) and brown trout. For the other tests that did not
have muscle or whole body selenium measurements, conversion factors (CFs) were used to
convert the egg/ovary CV to a muscle or whole body CV. The direct calculation of the muscle
and whole body CVs (when data were available) reduced uncertainty in these effect level
estimates.
6.3.1.4	C onversi on F actors
When muscle or whole body chronic values could not be determined directly using
selenium concentrations measured and reported for the respective tissue, conversion factors (CF)
were used to convert the egg/ovary chronic value to either a muscle or whole body chronic value.
To derive egg-ovary to whole-body CF values, EPA defined matched pairs of selenium
measurements from the eggs or ovaries and from the whole-body measured from the same
individual fish or from matched composite samples. If multiple measurements from both eggs
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and ovaries of the same individual or matched composite sample were available, the average
value was used. Similar pairings were done for egg-ovary to muscle CF values.
After the data sets of the pairings were compiled, EPA first confirmed a statistical
relationship between egg-ovary and whole body selenium for each species using ordinary least
squares (OLS) linear regression of the matched pairs of measurements. If the regression resulted
in a significant fit (P<0.05) with a positive slope, EPA calculated the ratio of the egg-ovary to
whole body (or muscle) selenium concentration of each matched pair and used the median ratio
as the CF value for the species. A detailed comparison of the advantages and disadvantages of
the median ratio and least squares regression approaches to calculating CFs, along with a
comparison of CFs calculated from median ratios, OLS regression following log transformation,
and total least squares (TLS) regression following log transformation is in Appendix N. Table N-
3 provides a comparison of the median-based and regression-based CFs when they are used to
convert an egg-ovary selenium concentration to muscle or whole body. Generally, the median-
based and TLS-based CFs were similar for both tissue types and this similarity resulted in
similar criterion element values (bottom row of Table N-3). The muscle criterion element value
for the data set that contained directly calculated CVs and converted CVs was similar whether
median or TLS CFs were used, 11.3 and 10.2, respectively. The whole body criterion element
value was also similar using these two approaches, 8.5 and 9.4, respectively. The median-based
CF approach was considered to be better than the regression-based CF approaches at reducing
uncertainty. A detailed comparison and rationale for the median approach is discussed in
appendix N.
EPA had sufficient egg-ovary and whole-body selenium measurements to directly
calculate egg-ovary to whole body CF values for 13 species of fish. Similarly, there were
sufficient egg-ovary and muscle selenium measurements to directly calculate egg-ovary to whole
body CF values for 16 species of fish. To derive CF values for additional fish species, EPA used
a taxonomic-relatedness approach (most similar taxon) approach to estimate CF. This approach
is consistent to that done for TTF estimates, and is described in Section 3.2.2, and in greater
detail in Appendix B.
The variability of CFs between fish species and within fish species was fairly low. EPA
derived 13 CF values directly from matched pairs of egg-ovary and whole-body selenium
measurements and an additional seven CF values by multiplying EO/M and M/WB conversion
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factors (Table 3.12). Excluding mountain whitefish (CF= 7.4), CFs for 19 of the 20 species
ranged from 1.20 to 3.11, a 2.6-fold difference. CF variability within each species was also low
for 11 of the 13 species for which egg-ovary to whole-body CFs were determined directly and a
standard deviation calculated (Table 3.12). The two species with relatively high standard
deviations contained data that were potentially anomalous. When the potentially anomalous data
were removed the standard deviations for these two species were reduced considerably (see
footnote to Table 3.12).
6.3.2 Trophic Transfer Factors
A Trophic Transfer Factor (TTF) represents the transfer of selenium from one trophic
level to the next higher trophic level. T'/T's are used in the translation of the tissue criterion
element concentration to a water element value. For a description of how TTFs are used in
translation, see Section 3.2.1, Translation from Fish Tissue Concentration to Water Column
Concentration.
Similar to CFs, EPA calculated TTFs from field data using the median-ratio approach
after first performing OLS regression of matched pairs of selenium measurements for the two
taxa representing successive trophic levels to determine if the relationship is significant (P<0.05)
and has a positive slope. EPA also evaluated using only OLS regression results to calculate TTF
values. OLS regression was performed using matched concentrations of selenium in the food of a
particular species or taxonomic group with the concentration of selenium in the organism's
tissue, and then the slope of the regression was used as the TTF for that species or taxonomic
group. An advantage of the regression approach is that it estimates the quantitative relationship
of selenium across a range of environmental concentrations in a manner that allows statistical
assessment. Disadvantages of this regression approach include the assumption that the
underlying data are normally distributed; the possibility that one or a few very high or low values
can have a disproportionate influence on the slope of the fitted line; and the fact that the
bioaccumulation model does not account for a non-zero y-intercept. Constraining the y-intercept
to zero (also known as regression through the origin or RTO) eliminates the added complexity of
a non-zero y-intercept. However, RTO further increases the disproportionate influence of one or
a few high values on the slope of the fitted line. Furthermore, RTO does not provide a
straightforward way of evaluating goodness of fit (Gordon 1981).
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The median-ratio approach, following confirmation of a significant (P<0.05) relationship
and positive slope, was considered to be more appropriate for deriving T'/T's from field data that
the OLS regression approach. Requiring a significant positive OLS linear regression coefficient
confirms the relationship between selenium in organisms and the food they ingest is adequately
represented by the available data. Using the median of the individual ratios provides an estimate
of central tendency for that relationship that is less sensitive to potential bias from measurements
taken from aquatic systems with very high or very low selenium concentrations. Some aquatic
organisms exhibit selenium bioaccumulation inversely related to water concentration (McGeer et
al. 2003; Borgman et al. 2004; DeForest et al. 2007). This inverse relationship is likely due to
saturation uptake kinetics of specific transport mechanisms that regulate metals bioaccumulation
within certain ranges (U.S. EPA 2007). EPA evaluated the effect of very high and very low
selenium concentrations on the calculation of TTF values using the hybrid approach (use of
median ratios for matched data with significant relationship and positive slope) described above
by excluding selenium measurements above various minimum and/or below various maximum
selenium concentrations. EPA found that using the median ratio effectively attenuates any effects
of selenium concentration on the calculation of TTF values using the hybrid approach described
above without the need to introduce additional arbitrary exclusion criteria.
TTFs were also determined using physiological coefficients (see Section 3.2.2.1,
Derivation of Trophic Transfer Factors (TFF) Values. However, if a TTF value could be
calculated from both physiological coefficients and field data, EPA used the TTF value
calculated from the substantially larger number of field measurements to minimize statistical
uncertainty.
TTFs were calculated for 32 fish species, and ranged from 0.68 to 2.67 (Table 3.11). The
majority of fish TTFs fell within a relatively narrow range, with an interquartile range (25th - 75th
centile) of 1.03 to 1.42. Variability of TTFs among the 13 invertebrate taxa was higher, ranging
from 0.74 to 4.58 (Table 3.10). Much of the variability among invertebrate TTFs was related to
taxonomic groups. The two bivalve TTFs ranged from 4.00 to 4.58. The five insect TTFs ranged
from 1.48 to 2.88, the five crustacean TTFs ranged from 0.74 to 1.89, and the TTF for
blackworms was 1.29.
EPA translated the tissue criterion element concentration to water element values at field
sites that had selenium measurements in the required water, particulates, invertebrates, and fish.
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For species without sufficient data to directly calculate a TTF value at these sites, EPA estimated
the TTF value by sequentially considering higher taxonomic classifications until one or more
taxa for which a calculated TTF value was available matched the taxon being considered. If the
lowest matching taxon was common to more than one species with a TTF value available, EPA
used the median TTF from the matching species.
6.3.3 Enrichment Factors
Enrichment factors are discussed in Section 3.2.2.3 and Appendix H. This factor,
describing how the bottom of the food chain takes up selenium, is the most variable between
sites. Variability among EFs is the main reason that fish BAFs vary so much between sites, and
this variability is the reason the national criterion for selenium needed to be tiered, with tissue
having priority over water, to increase certainty that the criterion is protective as intended. The
range of site EF values shown in Appendix H spans more than a 100-fold range.
The EF value measured at a particular site is also likely to be the site's most uncertain
parameter, being a ratio of measurements of algae, detritus, and sediment, which may vary
within a site in uncertain ways, and measurements of water, which vary over time. The approach
for setting site EFs was designed to reduce uncertainty. As described in Appendix H, EPA
calculated EF values by searching its database of selenium measurements and identifying all the
selenium measurements from algae, detritus, or sediment. EPA then searched for corresponding
water column measurements from samples collected at the same aquatic site within one year of
the particulate sample. If more than one water concentration was available for any given
particulate measurement, the median water concentration was used. For each of these matched
pairs of particulate and water measurements, EPA calculated the ratio of particulate
concentration to water concentration. If more than one ratio for any given category of particulate
material (algae, detritus, or sediment) was calculated at an aquatic site, EPA used the median of
those ratios. Selenium concentrations between particulate and water concentrations were higher
for algae and detritus than for sediment. To reduce uncertainty in EF values associated with
sediments, at least two particulate selenium measurements with corresponding water column
measurements were required, and sediment measurements were used only if there was at least
one other measurement from either algae or detritus. The geometric mean of the algae, detritus,
and sediment ratios was then calculated and used as the site EF. Because there were at most only
3 possible values (one for algae, one for detritus, and one for sediment), EPA used the geometric
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mean in order to reduce the potential for one of the values to have excessive influence on the
final site EF value. Sites with insufficient data to fulfill these data requirements were not used.
Had EPA increased the data requirements for setting a site EF, then the database would
be restricted to a smaller number of sites. Because the variability between sites is high, reducing
the number of sites in the database would decrease the confidence in the representativeness of
the few sites retained (that is, it would increase the potential for sampling error in attempting to
characterize the nation's waters). The EF determination process thus involved a balance between
having enough information to reasonably characterize each site, and having enough sites to
represent the range of the nation's waters.
Inclusion of selenium speciation information (such as selenate and selenite
concentrations) was infeasible. Very few sites would have the requisite information, thereby
increasing the uncertainty in the representativeness of any possible derived national criterion.
Likewise, inclusion of a sulfate relationship was not feasible on a national basis at this time, for
lack of sulfate data at many sites in the database.
Because EF is the BAF component that varies the most between sites, it is the most
important in determining what the water concentration would be for a site if its fish tissue
concentrations were hypothesized to be at the level of the fish tissue criterion element value.
That is, sites with higher EFs tabulated in Appendix H have lower translated water
concentrations in Table 3.13. Despite uncertainties in this parameter, model-predicted versus
observed fish-tissue concentrations in the vicinity of the water element criterion concentrations
are relatively unbiased, as shown in the figures of Appendix I.
For particular sites, the appropriateness of the national criterion can be resolved by site
specific criteria when necessary (e.g., when a permit limit for water is required), as
recommended in Appendix K. When taking measurements of a site, uncertainty in particulate
measurements (the numerator of the EF) can be bypassed by using site-specific fish BAFs, since
they only consider water and fish tissue selenium measurements. On the other hand, uncertainty
in characterizing time-variable water concentrations is a problem shared by EFs and BAFs.
However, this uncertainty can be reduced by sampling in a spatially and temporally robust
manner, appropriate for the site in question, and then using the mathematical modeling approach
to derive a site specific criterion.
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6.3.4 Water Values
Derivation of the water criterion element from the egg-ovary criterion element is
described in Sections 2.7.8 and 3.2, and involves EFs, TTFs, and CFs. Uncertainties in predicted
tissue-to-water ratios combine the uncertainties in the parameters from which they are predicted.
The prediction model is linear in all respects. Potential nonlinearities are therefore an
uncertainty. Section 6.2.1 and Appendix I assesses the accuracy of the predictions. As shown in
the figures of Appendix I, the predicted values perform reasonably well in the vicinity of the
water criterion element concentrations.
Although an earlier published draft document weighted sites by the number of fish
species sampled (between 1 to 6 species per site), that overweighting of sites with several
measured species was removed from this draft by using only the most bioaccumulative fish
species per site, thereby reducing uncertainty that the fish tissue criterion element will be
exceeded when using only water column concentration data. Lentic and lotic sites were assessed
separately, per Section 3.2.4, This increases the likelihood that the water criterion element
concentration will be appropriate for the site of application.
To reduce the likelihood that the water criterion element concentration will be under-
protective for any particular site of application, the 20th percentiles of translated water
concentrations for all lentic and lotic sites, respectively, were used as the water criterion element
concentrations. As described previously, these distributions represented the translated water
concentrations for the most bioaccumulative fish species at each site, which further reduces the
likelihood that the fish tissue criterion element would be exceeded if the water criterion element
was being met. These water criterion elements should not be interpreted to be potentially under-
protective in 20 percent of sites, because when applied to a site's 30-day once-in-three-year
maximum concentration (which is higher than its median), the 20th percentile site would not
attain. The actual percentage of sites protected would thus be greater than 80 percent, but the
exact percentage is uncertain.
6.4 Protection of Threatened or Endangered Species
The chronic toxicity dataset for selenium contains toxicity data for two Federally-listed
endangered species, Cyprinodon macularius (desert pupfish) and Oncorhynchus mykiss (listed as
steelhead, indicating anadromous individuals, but herein called rainbow trout, implying non-
anadromous individuals). The dataset also contains toxicity data for Acipenser transmontanus
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(white sturgeon), which is listed as endangered in specific locations, such as the Kootenai River
white sturgeon in Idaho and Montana. The white sturgeon also serves as a surrogate for other
sturgeon listed as threatened or endangered (e.g., pallid and shovelnose sturgeon). The Acipenser
GMCV of 15.6 mg/kg dw egg is the lowest value in the dataset and therefore provides protection
for other potentially sensitive sturgeon. The white sturgeon chronic value is greater than the
chronic egg-ovary criterion element value.
Desert pupfish, Cyprinodon macularius, with a chronic value estimated to be >27 mg
Se/kg dw egg, is not among the most sensitive species. Its chronic value of >27 mg Se/kg dw egg
is substantially above the chronic egg-ovary criterion element value of 15.1 mg Se/kg dw.
Oncorhynchus mykiss has an SMCV of 24.5 mg Se/kg dw egg, whose genus is the fourth
most sensitive species in the dataset. The dataset contains multiple studies with cutthroat trout
(Oncorhynchus clarkii) some subspecies of which are Federally listed as threatened. The SMCV
for cutthroat trout is 26.2 mg Se/kg dw egg. Both of these chronic values for Oncorhynchus
species are greater than the chronic egg-ovary criterion element.
The dataset also contains toxicity information for Salvelinus malma (Dolly Varden)
which is not threatened or endangered, but is so closely related to the threatened Salvelinus
confluentus (bull trout) that it can hybridize with that species, producing fertile offspring (Baxter
et al. 1997). Dolly Varden is the least sensitive fish species for which information is available,
with an SMCV of 56 mg Se/kg dw egg. Salvelinus fontinalis, brook trout, can also hybridize
with bull trout, but the offspring are sterile, suggesting that it is less closely related. With the
available study of brook trout, although in Section 6.1.5 the NOEC is conservatively set to >20.5
mg Se/kg dw egg, which was the average concentration at the Holm et al. (2005) high-exposure
site. The concentration-response information for the offspring of individual females, presented in
Appendix C, suggests that its ECio could be substantially higher, possibly as high as that for
Dolly Varden.
The egg-ovary criterion element value of 15.1 mg Se/kg (dw) is below all of the above
mentioned chronic egg-ovary values for threatened and endangered (or closely related) species.
However, because other threatened or endangered species could be more sensitive, if relevant
new information becomes available in the future, it should be considered in state- or site-specific
criterion calculations.
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The protectiveness of the whole body criterion element concertation of 8.5 mg/kg dw to
threatened and endangered species is also supported by a recent non-reproductive study with two
sturgeon species. De Riu et al. (2014) fed juvenile green and white sturgeon (-30 g body weight)
diets containing a range of selenium concentrations (selenomethionine added to diet formulation;
2.2	mg/kg Se in control diet (no added Se) and 19.7, 40.1 and 77.7 mg/kg Se in the three
treatment diets). Several endpoints were monitored over the 8-week exposure period including
survival and percent body weight increase (% BWI). White sturgeon had no mortalities through
the highest dietary treatment. Green sturgeon juveniles had 0%, 7.7% and 23.1% mortality with
the three dietary treatments. TRAP analysis (threshold sigmoid nonlinear regression) of the green
sturgeon survival data resulted in a whole body ECio value of 28.93 mg/kg dw. ECio values were
lower for % BWI using TRAP. For % BWI, the whole body ECio value for green sturgeon was
16.36 mg/kg dw, and 23.94 mg/kg dw for white sturgeon.
Also notable, the background concentrations of selenium in the juvenile green and white
sturgeon were also elevated at 7.2, 6.5 and 7.1 mg/kg dw (green sturgeon whole body), and 4.8
7.3	and 5.6 mg/kg dw (white sturgeon whole body) at test initiation, and after four and eight
weeks of exposure, respectively.
The De Riu et al. (2014) study suggests that green sturgeon may be more sensitive to
selenium than white sturgeon and also that the EPA whole body concentration of 8.5 mg/kg dw
will be protective, based on the survival and growth data and the observation in De Riu 2014 that
the control whole body tissue concentrations (up to 7.2 mg/kg dw) are approaching the proposed
criterion. This is important because white sturgeon, as well as juvenile green sturgeon (up to
three to four years), spend most of their time in the coastal rivers and estuaries. All species in the
Acipenseriformes (sturgeon and paddlefish) spawn in freshwaters (Bemis and Kynard 1997) or
spend their entire life in freshwater. The white sturgeon's ECio in the dataset provides surrogacy
for the threatened and endangered species from this group. For more information on the De Riu
et al. (2014) study, see Appendix E.
6.4.1 Special Consideration for Pacific Salmonid Juveniles
The current criterion is based on reproductive effects (larval mortality and/or deformities)
for offspring of selenium-exposed adults, and the whole-body criterion element is derived from
the egg-ovary element, with an implicit assumption of adult exposure to selenium. One peer-
reviewer of the 2014 EPA External Peer Review Draft criterion document raised concerns
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regarding the protection of anadromous salmonids, since there is at least some evidence (e.g.,
Hamilton et al. 1990) that juvenile growth may be comparable in sensitivity to reproductive
effects endpoints used by EPA. Anadromous salmon species (e.g., Chinook salmon) in the
Pacific Northwest are unique in that reproductively mature adults are not exposed to selenium in
the freshwater environment due to their life history; young juvenile salmon leave freshwater
streams and rivers as smolts and mature to adulthood in the marine environment until migration
for spawning begins. Furthermore, they are semelparous, breeding only once in their lifetime and
subsequently dying, so there is no potential selenium exposure following spawning in freshwater.
Juvenile salmon have evolved different strategies for growth and maturation to the smolt
stage, and may spend from three months to two years in freshwater (depending on timing of egg
hatching and other factors) before migrating to estuarine areas as smolts and into the ocean to
feed and mature. Salmon remain in the ocean for one to six years (more commonly two to four
years), with the exception of a small proportion of yearling males (called jack salmon), which
mature in freshwater or return after two or three months in salt water (NOAA 2011).
The physiological and morphological changes that allow these species to adapt to marine
conditions as juveniles are reversed in returning adults preparing to migrate up natal streams to
spawn. One key change is the cessation of feeding prior to re-entry into freshwater. Since mature
females are not feeding after returning to freshwater, it is not representative to predict
reproductive effects for anadromous salmonid species based on egg-ovary selenium
concentrations, because the exposure is wholly from selenium sources in the marine environment
(Groot and Margolis 1991).
6.4.1.1 Selenium Toxicity to Juvenile Salmonids
Hamilton et al. (1990) assessed the toxicity of two organoselenium diets in 90-day partial
life cycle tests in freshwater with two life stages of Chinook salmon (Oncorhynchus
tshawytscha). The first diet consisted of fish meal made from low-selenium mosquitofish
(collected from a reference site) fortified with selenomethionine (here termed the SeMet diet).
The second diet contained fish meal made from high-selenium mosquitofish (Gambusia affinis)
collected from the San Luis Drain (SLD), California (here termed the SLD diet). This waterbody
is known to have high concentrations of selenium. A 90-day partial life cycle study was
conducted with swim-up stage salmon larvae in a standardized fresh water that simulated
dilution of San Luis Drain water.
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Survival and growth (length and weight) were measured at 30, 60, and 90 days.
Unexplained control mortality (33%) between day 60 and day 90 introduced an unacceptable
level of uncertainty into the overall health of the fish. The 1985 Aquatic Life Guidelines
(Stephan et al. 1985) and the Manual of Instructions for Preparing Aquatic Life Water Quality
Criteria Documents (Stephan 1987) require that excessive control mortality be treated as an
exclusionary threshold in data quality assessments for regulatory purposes such as deriving water
quality criteria. Therefore the 90-day survival data from this study was not used quantitatively.
At 60 days, larval control mortality was acceptable (1%), and 60-day larval survival was > 90%
in all SLD and SeMet treatments (3.2 ppm - 18.2 ppm) except for the high Se treatment (35.4
ppm). Whole body selenium concentrations were measured at 60 days, were 10.4 and 13.3 mg/kg
dw, respectively, for larvae fed the SeMet and SLD diets of 18.2 mg/kg dw (Hamilton et al.
1990).
Although survival was similar in response to the two diets, larval growth responses
differed between the SLD and SeMet diets. The salmon fed the SeMet mosquitofish diet had
significant reductions in both length and weight at 30, 60, and 90 days; but only at the two
highest concentrations (18.2 and 35.4 ppm). The average length and weight of the larvae fed the
SLD mosquitofish diet were significantly lower at all concentrations at 30, 60, and 90 days. The
greater effect on growth parameters fed the SLD mosquitofish meal diet could have been caused
by one or more of several factors: 1) additional forms of organic selenium (e.g., selenocysteine)
present in the SLD mosquitofish, 2) additional toxic elements (e.g., heavy metals) that were
accumulated by the SLD mosquitofish, and not present in the reference site mosquitofish, and 3)
differential metabolic processing of the organoselenium contained in the proteins of the SLD
mosquitofish and fed to the larval salmon, versus the larvae fed the diet containing the free
amino acid selenomethionine (Hamilton et.al. 1990).
EPA performed a regression on the 60-day weight and whole body concentrations, and
derived a whole body ECio value of 7.355 mg/kg dw for the SeMet diet for reduced growth, and
a whole body ECio value of 11.14 |ig/g dw for the SLD diet for reduced growth. These values
are the only two available ECio Species Mean Chronic Values (SMCVs) for non-reproductive
endpoints for the genus Oncorhynchus, and the Genus Mean Chronic Value (GMCV) is 9.052
mg/kg dw. This is greater than the national whole body criterion element concentration of 8.5
mg/kg dw, which will thus be protective of this genus.
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EPA recommends that states and tribes consider use of the whole-body criterion element
for juvenile (smolt) anadromous Pacific salmon species as the primary criterion element over the
other elements due to the unique life history of these species, specifically, the lack of exposure to
adult salmonids from selenium in freshwater prior to reproduction. The hierarchal structure of
the egg-ovary tissue over the other tissue criterion elements applies to all other species in the
family Salmonidae. The egg-ovary criterion element, as well as the other fish tissue criterion
elements and the water column criterion elements still apply, as applicable, to protect the
remainder of the aquatic community in these waters.
6.5 Aquatic-Dependent Wildlife is Beyond the Scope of this Aquatic
Criterion Derivation
AWQC that are developed by EPA typically focus directly on aquatic life, not aquatic-
dependent wildlife such as birds. As presented by Campbell (2011), EPA recognizes that
selenium effects on aquatic-dependent wildlife are also of concern but considers them beyond
the scope of this national criterion update. In the interest of providing updated guidance to
protect against the known risks of selenium exposure to fish, EPA decided to focus its analyses
on updating the existing selenium criterion for freshwater aquatic life based on the latest
scientific evidence.
In the future, EPA plans to consider the effects of selenium on aquatic-dependent
wildlife, potentially in the form of criteria expanded to address aquatic-dependent wildlife. When
translated to a water concentration, a criterion protective of aquatic-dependent wildlife may be
more stringent or less stringent than the values provided for aquatic life in this criterion
document. This is because data indicate that for most ecosystems, selenium concentrations are
generally conserved or increase incrementally at each trophic level in a food web (after a
substantial increase from water to trophic level 1 (e.g., algae). Certain specific ecosystems (e.g.,
estuarine and marine systems more commonly) with mollusk-based food-webs may create a
pathway for more selenium to bioaccumulate, particularly in molluscivorous predators (certain
fish and aquatic bird species), since the available data indicate that mollusks generally have a
higher trophic transfer factor than other invertebrate taxa. This level of bioaccumulation is
typically lower, and in contrast to other bioaccumulative chemicals such as mercury, which have
much greater biomagnification.
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As stated previously, the single largest step in tissue selenium accumulation in aquatic
environments occurs at the base of the food web where algae and other microorganisms
accumulate selenium from water (Orr et al. 2012; Stewart et al. 2010). Mollusks such as mussels
and clams accumulate selenium to a much greater extent than planktonic crustaceans and insects
due to higher ingestion rates of both particulate-bound (algae) and dissolved selenium from the
water column through filter feeding, and these organisms have a lower selenium elimination rate
(Luoma and Rainbow 2005). Thus, aquatic-dependent wildlife criteria for species that are
primarily molluscivores may have concentrations of concern that are not protected by the 2016
selenium criterion elements found in this document. The criteria values for aquatic-dependent
wildlife would be expected to depend on the aquatic systems, species, and food webs considered,
as well as spatial and temporal considerations related to selenium exposure and breeding and
nesting seasons. Where sensitive aquatic-dependent (e.g., bird) species are known to exist, states
should consider developing site-specific criteria based on data for such species.
6.6 Summary
EPA developed the 2016 national 304(a) Aquatic Life Ambient Water Quality Criterion
for Selenium in Freshwater to be protective of most aquatic life genera in most waters of the
United States, with an intended goal of protecting approximately 95% of aquatic genera in an
ecosystem. This freshwater chronic selenium criterion applies only to aquatic life, and is not
intended to address selenium toxicity to aquatic-dependent wildlife such as aquatic-dependent
birds. This document provides guidance to States and Tribes authorized to adopt water quality
standards under the Clean Water Act (CWA), to protect aquatic life from toxic effects of
selenium.
The 2016 selenium criterion is a chronic criterion that is composed of four elements. All
elements are protective against chronic selenium effects. Two elements are based on the
concentration of selenium in fish tissue and two elements are based on the concentration of
selenium in the water-column. The recommended elements are: (1) a fish egg-ovary element; (2)
a fish whole-body and/or muscle element; (3) a water column - element (one value for lentic and
one value for lotic aquatic systems); and (4) a water column intermittent element to account for
potential chronic effects from short-term exposures (one value for lentic and one value for lotic
aquatic systems). The assessment of the available data for fish, invertebrates, and amphibians
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indicates that a criterion value derived from fish is expected to be protective of the aquatic
community, based on available data.
EPA recommends that states and tribes adopt into their water quality standards a
selenium criterion that includes all four elements, and express the four elements as a single
criterion composed of multiple parts, in a manner that explicitly affirms that the whole-body or
muscle elements supersede the water column element, and the egg-ovary element supersedes any
other element. The magnitude of the fish egg-ovary element is derived from analysis of the
available toxicity data. The magnitudes of the fish whole-body element and fish muscle elements
are derived from the egg-ovary element coupled with data on concentration ratios among tissues.
The magnitudes of the water column elements are derived from the egg-ovary element coupled
with bioaccumulation considerations. Inclusion of the fish whole-body or fish muscle element
into the selenium criterion ensures the protection of aquatic life when fish egg or ovary tissue
measurements are not available, and inclusion of the water column elements into the selenium
criterion ensures protection when neither fish egg-ovary nor fish whole-body nor muscle tissue
measurements are available. There are two specific circumstances where the fish tissue
concentrations do not fully represent potential adverse effects on fish and the aquatic ecosystem:
1) "fishless" waters, because of the inability to collect sufficient fish tissue to measure selenium
concentrations in fish in such waters, and 2) areas with new selenium inputs, because the fish
tissue concentrations in such systems would not yet represent steady state conditions upon which
the criterion is based.
To ensure that the contribution of short-term exposures to the bioaccumulation risks is
accounted for in all situations, EPA is recommending that the intermittent exposure element be
included in the selenium criterion, as noted above. EPA is not recommending a separate acute
criterion element derived from the results of toxicity tests having water-only exposure because
selenium is bioaccumulative and toxicity primarily occurs through dietary exposure. Application
of the intermittent exposure criterion element values to single day, high exposure events will
provide protection from the most important selenium toxicity effect, reproductive toxicity, by
protecting against selenium bioaccumulation in the aquatic ecosystem resulting from short-term,
high exposure events.
The egg/ovary-based tissue criterion element of 15.1 mg Se/kg dw is based on a genus
sensitivity distribution that used the most sensitive assessment endpoint observed in toxicity
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tests, reproductive effects, and included fish species known to be sensitive to selenium (i.e.,
species from Salmonidae and Centrarchidae), as well as three endangered species (desert
pupfish, rainbow trout and white sturgeon).
With respect to the chronic water column criterion elements, EPA intends the lentic and
lotic values of 1.5 and 3.1 |ig/L, respectively, to be protective of most surface waters in the U.S.
These water concentrations represent the 20th percentile of the distribution of translated water
column values from sites across the U.S. The intermittent exposure water column criterion
element is derived from the chronic water column criterion element, which was derived from the
tissue-based criterion.
EPA recognizes selenium bioaccumulation potential depends on the structure of the food
web and several biogeochemical factors that characterize a particular aquatic system. Uncertainty
in the translation of the egg-ovary criterion element to the water column element can be reduced
by deriving a site-specific criterion that uses site-specific selenium data and information on food-
web dynamics from a biological assessment of the aquatic system. Appendix K provides
recommendations and examples for developing site-specific selenium criteria.
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Aquatic Life Ambient Water Quality
Criterion for
Selenium - Freshwater
2016
(Appendices A-N)
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, D.C.

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List of Appendices
APPENDIX A: Selenium Chemistry	A-l
1.0 Inorganic Selenium	A-2
2.0 Organoselenium	A-4
3.0 Departure from Thermodynamic Equilibrium	A-5
4.0 Physical Distribution of Species in Surface Water	A-5
APPENDIX B: Conversions	B-1
1.0 Conversion of Wet to Dry Tissue Weight	B-2
1.1	Methodology	B-2
2.0 Derivation of Tissue Conversion Factors	B-5
2.2	CF values calculated directly from whole-body and egg-ovary selenium
measurements	B-7
2.3	Muscle to egg-ovary conversion factors	B-22
2.4	Muscle to whole-body conversion factors	B-42
3.0 Derivation of Trophic Transfer Factor Values	B-59
3.1	Methodology	B-59
3.2	TTF values from physiological coefficients	B-61
3.2.1	Invertebrates	B-61
3.2.2	Vertebrates	B-66
3.3	TTF values from field data	B-68
3.3.1	Invertebrates	B-68
3.3.2	Vertebrates	B-81
4.0 Food Web Models Used to Calculate Composite TTFs to Translate the Egg-Ovary
FCV to Water-Column Values	B-163
APPENDIX C: Summaries of Chronic Studies Considered for Criteria Derivation	C-l
APPENDIX D: Summary Studies of Non-Reproductive Effects	D-l
1.0 Studies of Non-Reproductive Effects	D-2
1.1	Acipenseridae	D-2
1.1.1 Acipenser transmontanus (white sturgeon)	D-2
1.2	Cyprinidae	D-2
1.2.1	Pogonichthys macrolepidotus (Sacramento splittail)	D-2
1.2.2	Pimephales promelas (fathead minnows)	D-2
1.3	Catostomidae	D-4
1.3.1	Xyrauchen texanus (razorback sucker)	D-4
1.3.2	Catostomus latipinnis (flannelmouth sucker)	D-5
1.4	Salmonidae	D-5
1.4.1	Oncorhynchus tshawytscha (Chinook salmon)	D-5
1.4.2	Oncorhynchus mykiss (rainbow trout)	D-6
1.5	Moronidae	D-6
li

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1.5.1 Morone saxitilis (striped bass)	D-6
1.6 Centrarchidae	D-7
1.6.1 Lepomis macrochirus (bluegill)	D-7
APPENDIX E: Other Data	E-1
1.0 Selenite	E-2
2.0 Selenate	E-2
3.0 Other Data - Endangered Species	E-13
4.0 Other Data - Chronic Studies with Fish Species	E-21
4.1 Evaluation of zebrafish (Danio rerio) and native cyprinid sensitivity to selenium...E-41
4.1.1	Part I. Chronic summary of Thomas (2014) and Thomas and
Janz (2014)	I >43
4.1.2	Part II - Evaluating Sensitivity of Cyprinids (Cyprinidae) to Selenium
from Field and Laboratory Data	E-46
5.0 Other Data - Chronic Studies with Invertebrate Species	E-58
5.1	Rotifers	E-58
5.2	Aquatic Worms	E-58
5.3	Aquatic Insects (Plecoptera: Mayfly)	E-58
5.4	Aquatic Insect (Midge: Chironimids)	E-62
6.0 Other Data - Field Study West Virginia Impoundments	E-62
7.0 Other Data - Nutritional Deficiency/Sufficiency Studies Containing Measured
Selenium in the Diet and Whole Body Fish Tissue	E-64
APPENDIX F: Toxi city of S el enium to Aquati c PI ants	F-l
1.0 Selenite	F-2
2.0 Selenate	F-2
APPENDIX G: Unused Data	G-1
APPENDIXH: Calculation of EF Values	II-l
APPENDIX I: Observed Versus Predicted Egg-Ovary Concentrations	1-1
APPENDIX J: Supplementary Information on Selenium Bioaccumulation in
Aquatic Animals	J-l
1.0 Effects of Growth Rate on the Accumulation of Selenium in Fish	J-2
2.0 Analysis of the Relative Contribution of Aqueous and Dietary Uptake on the
Bioaccumulation of Selenium	J-3
3.0 Kinetics of Accumulation and Depuration: Averaging Period	J-3
3.1	Background	J-3
3.2	Approach for Modeling Effects of Time-Variable Se Concentrations	J-5
3.2.1	Model Results	J-8
3.2.2	Summary of Scenario Results	J-10
3.2.3	Example Responses to Increases in Water Concentrations	J-l 1
APPENDIX K: Translation of a Selenium Fish Tissue Criterion Element to a
Site-Specific Water Column Value	K-l
in

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1.0 Translating the Concentration of Selenium in Tissue to a Concentration in Water
Using Mechanistic Bioaccumulation Modeling	K-2
1.1	Relating the Concentration of Selenium in Fish Tissue and Water using the
Mechanistic Modeling Approach	K-4
1.2	Steps for Deriving a Site-Specific Water Concentration Value from the
Egg-Ovary Criterion Element	K-9
1.2.1	Identify the Appropriate Target Fish Species	K-9
1.2.2	Model the Food-Web of the Targeted Fish Species	K-13
1.2.3	Identify Appropriate TTF Values	K-14
1.2.4	Determine the Appropriate EF Value	K-18
1.2.5	Determine the Appropriate CF Value	K-21
1.2.6	Translate the Selenium Egg-Ovary Criterion Element into a
Site-Specific Water Concentration Value using Equation K-l	K-24
1.3	Managing Uncertainty using the Mechanistic Modeling Approach	K-24
1.4	Example Calculations	K-25
1.4.1	Example 1	K-26
1.4.2	Example 2	K-27
1.4.3	Example 3	K-28
1.4.4	Example 4	K-29
1.5.5	Example 5	K-30
1.5.6	Example 6	K-31
2.0 Translating the Concentration of Selenium in Tissue to a Concentration in Water
using Bioaccumulation Factors (BAF)	K-32
2.1	Summary of the BAF Approach	K-32
2.1.1 Example: Derivation of a site specific water column criterion for a
waterbody impacted by selenium	K-33
2.2	Managing Uncertainty using the BAF Approach	K-33
3.0 Comparison of Mechanistic Bioaccumulation Modeling and BAF Approaches	K-36
3.1	Translation using the BAF Approach	K-39
3.2	Translation using the Mechanistic Bioaccumulation Modeling Approach	K-40
3.3	Summary Comparison of the Mechanistic Bioaccumulation and BAF
Approaches	K-43
APPENDIX L: Analytical Methods for Measuring Selenium	L-l
1.0 General Considerations when Measuring Concentrations of Selenium	L-2
2.0 Analytical Methods Recommended for Measuring Selenium in Water	L-2
2.1	American Public Health Standard Method 3114 B	L-3
2.2	EPA Method 200.8	L-4
2.3	EPA Method 200.9	L-4
3.0 Analytical Methods Available for Measuring Selenium in Fish Tissue	L-5
3.1 Strong acid digestion	L-5
IV

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3.2	Dry-ashing digestion	L-6
3.3	Analytical methods available for measuring selenium in particulate material	L-6
APPENDIX M: Abbreviations	M-l
Reference and site abbreviations	M-2
APPENDIX N: Comparison of Approaches for Calculating Selenium Tissue
Conversion Factors	N-l
1.0 Comparison of the Median Ratio and Regression Approaches	N-2
2.0 Comparison of the Ordinary Least Squares and Total Least Squares Regression
Approaches	N-6
2.1	Example 1 - Flannelmouth Sucker (Egg-ovary/Whole-body)	N-9
2.2	Example 2 - Bluegill (Egg-ovary/Whole-body)	N-12
3.0 Comparison of Median- and Regression-based Conversion Factors to Calculate
Chronic Values for Muscle and Whole Body Tissues	N-14
v

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APPENDIX A: Selenium Chemistry
A-l

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Selenium in aquatic ecosystems exists in a broad range of oxidation states: (+ VI) in selenates
(HSe04, Se042") and selenic acid (H2Se04), (+ IV) in selenites (HSe03, Se032") and selenous acid
(H2Se03), 0 in elemental selenium, and (-II) in selenides (Se2~, HSe), hydrogen selenide (H2Se), and
organic selenides (R2Se). Selenium also shows some tendency to form catenated species like organic
diselenides (RseSeR). Within the normal physiological pH range and the reduction potential range
permitted by water, only Se, Se032, HSe03, and Se042" can exist at thermodynamic equilibrium (Milne
1998). While ionic reactions are expected to be rapid in water, oxidation-reduction reactions may be slow,
and the possibility exists for the formation of HSe" in living systems and some environments where
anoxic conditions arise. The parallel behavior of comparable species of sulfur and selenium in living
systems has often been observed, but it is important to recognize that their chemical characteristics are
different in many ways. For instance, selenate is comparable to chromate in oxidizing strength and far
stronger than sulfate [(Se042/H2Se03) = 1.15 V; (Cr2072VCr3+) = 1.33V; (S0427H2S03) = 0.200V
(standard potentials in acid solution: Weast 1969)], whereas selenide is a much stronger reducing agent
than sulfide [(Se/H2Se) = -0.36 V; [S/H2S ]= 0.14V)].
1.0 Inorganic Selenium	
Selenate usually predominates in well-aerated surface waters, especially those with alkaline
conditions. In spite of its oxidizing strength, selenate (Se042 ) exhibits considerable kinetic stability in the
presence of reducing agents (Cotton and Wilkinson 1988). The radius of Se042" is comparable to that of
S042" (Frausto da Silva and Williams 1991), and uptake by cells is expected to take place via the same ion
channels or permeases for both anions. Competition between sulfate and selenate uptake has been
observed in many species: algae (Riedel and Sanders 1996), aquatic plants (Bailey et al. 1995),
crustaceans (Ogle and Knight 1996), fungi (Gharieb et al. 1995), HeLa cells (Yan and Frenkel 1994), and
wheat (Richter and Bergmann 1993). Reduced selenate bioconcentration with increasing sulfate
concentration has been demonstrated in Daphnia magna (Hansen et al. 1993). A significant inverse
relationship was shown to exist between acute selenate toxicity to aquatic organisms and ambient sulfate
concentrations (Brix et al. 2001a). Competition with selenate has also been observed for phosphate in
green algae (Riedel and Sanders 1996), and with chromate and tungstate in anaerobic bacteria (Oremland
et al. 1989).
Selenous acid species (HSe03" and Se032") can predominate in solution under the moderately
oxidizing conditions encountered in oxygenated waters. Between pH 3.5 and 9.0 biselenite ion is the
predominant ion in water, and at pH values below 7.0, selenites are rapidly reduced to elemental selenium
under mildly reducing conditions (Faust 1981), situations that are common in bottom sediments.
A-2

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Most selenite salts are less soluble than the corresponding selenates. The extremely low solubility
of ferric selenite Fe2(Se03)3 (Ks= 2.0  1.7 x 10"31), and of the basic ferric selenite Fe2(0H)4Se03 (Ks =
10"61 7), is important to the environmental cycling of selenium. Selenites also form stable adsorption
complexes with ferric oxides, forming complexes of even lower solubility than the ferric selenites. Under
certain conditions, selenite (in contrast to selenate) seems to be completely adsorbed in high amounts by
ferric hydroxide and, to a lesser extent, by aluminum hydroxide (Faust 1981). Coprecipitation techniques
have been applied for preconcentration of selenium in natural waters, using iron (III) hydroxides, which
coprecipitates selectively the selenite, but not the selenate, species in river and sea waters (Yoshii et al.
1977). Alum and iron coagulation precipitation can be used in water treatment processes to remove
selenite (Clifford et al. 1986). The low levels of selenium in ocean waters have been attributed to the
adsorption of selenite by the oxides of metals, such as iron and manganese (National Academy of
Sciences 1976).
Relative to selenate, selenite is more reactive because of its polar character, resulting from the
asymmetric electron density of the ion, its basicity (attraction to bond with proton), and its nucleophilicity
(attraction to bond to a nucleus using the lone pair electrons of the ion). No evidence has yet been
presented to show that HSe03" or Se032" is taken up intact into the cell interior. Evidence indicates that
selenite is reduced rapidly, even before uptake in some cases, making it difficult to distinguish between
uptake and metabolic processes (Milne 1998). Freshwater phytoplankton process selenate and selenite by
different mechanisms, leading to different concentrations within the cell, and the concentrations attained
are affected by various chemical and biological factors in the environment (Riedel et al. 1991). These
authors suggested that selenate is transported into the cell by a biological process with low affinity,
whereas selenite appears to be largely physically adsorbed. Contradictory evidence suggesting that
selenite uptake is enzymatically mediated was found with marine phytoplankton (Baines and Fisher
2001). Experimental results supporting the hypothesis that separate accumulation mechanisms for
selenate and selenite are present in D. magna have been published (Maier et al. 1993). However, while
some organisms appear to absorb selenite nonspecifically, specific transport systems exist in other
species. Sulfate competition is insignificant in the aquatic plant Ruppia maritima (Bailey et al. 1995), and
specific uptake systems have been demonstrated in some soft line microorganisms (Heider and Boeck
1993). Selenite uptake in green algae, unlike selenate, is increased substantially at lower pH values, a
property that represents another difference between these two anions (Riedel and Sanders 1996). The
uptake of inorganic selenium species, selenate and selenite, by the green alga Chlamydomonas reinhardtii
(Dang) was examined as a function of pH over the range 5 to 9, and in media with varying concentrations
of major ions and nutrients using 75Se as a radiotracer. Little difference was noted in the uptake of
selenate as a function of pH, with the maximum uptake found at pH 8; however, selenite uptake increased
A-3

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substantially at the lower pH values. Differences in speciation are suggested to be the cause of these
differences. Selenate exists as the divalent ion Se042" over the range of pH tested; whereas monovalent
biselenite ion HSe03" is prevalent at these pH values. At the low end of the pH range, neutral selenous
acid may also play a role.
Elemental selenium is not measurably soluble in water. It has been reported that elemental
selenium is slowly metabolized by several bacteria (Bacon and Ingledew 1989), and the translocation of
elemental selenium into the soft tissue of the marine mollusk Macoma balthica has been reported (Luoma
et al. 1992). The bioavailability of elemental selenium toM balthica was assessed by feeding the
organisms 75Se-labeled sediments in which the elemental selenium was precipitated by microbial
dissimilatory reduction. A 22% absorption efficiency of particulate elemental selenium was observed. In
view of the insolubility of elemental selenium, uptake may be preceded by air oxidation, or in reducing
environments thiols may facilitate the solubilization (Amaratunga and Milne 1994). Elemental selenium
can be the dominant fraction in sediments (Zawislanski and McGrath 1998).
Selenium is reduced to hydrogen selenide, H2Se, or other selenides at relatively low redox
potentials. Hydrogen selenide by itself is not expected to exist in the aquatic environment since the
Se/H2Se couple falls even below the H /H2 couple. Aqueous solutions of H2Se are actually unstable in air
due to its decomposition into elemental selenium and water. Under moderately reducing conditions,
heavy metals are precipitated as the selenides, which have extremely low solubilities. The following are
log Ks values of some heavy metal selenides of environmental interest: -11.5 (Mn2+), -26.0 (Fe2+), -60.8
(Cu+), -48.1 (Cu2+), -29.4 (Zn2+), -35.2 (Cd2+), and -64.5 (Hg2+). The precipitation of selenium as heavy
metal selenides can be an important factor affecting the cycling of the element in soils and natural waters.
2.0 Organoselenium	
Organic selenides (conventionally treated as Se(-II) species) in variable concentrations, usually in
the form of free and combined selenomethionine and selenocysteine, are also present in natural surface
waters (Fisher and Reinfelder 1991). Dissolved organic selenides may be an important source of selenium
for phytoplankton cells, because they can account for -80% of the dissolved selenium in open ocean
surface waters, and for a significant fraction in many other environments as well (Cutter 1989; Cutter and
Cutter 1995). Dissolved organoselenium levels of 14.2%, 65% and 66% were measured in samples (one
meter depth) from Hyco Reservoir, NC; Robinson Impoundment, SC; and Catfish Lake, NC; respectively
(Cutter 1986). The Hyco Reservoir organoselenium was identified as being protein bound.
Organoselenium concentrations were found to range from 10.4% (58.7 |ig/L) to 53.7% (1.02
|ig/L) of the total selenium present in Lake Creek and Benton Lake, MT surface waters (Zhang and
Moore 1996). Organoselenium quite often is measured as the difference between total dissolved selenium
A-4

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and the sum of selenite plus selenate, and is therefore not typically characterized. Much more work is
needed in the area of specific identification and characterization of the nature of the organic selenides
present in aquatic ecosystems. Organoselenium form(s) are much more bioavailable and probably play a
very important role in selenium ecotoxic effects (e.g. Besser et al., 1993; Rosetta and Knight 1995).
3.0 Departure from Thermodynamic Equilibrium	
In the highly dynamic natural waters, there is often a departure from thermodynamic equilibrium.
In the thermodynamic models, kinetic barriers to equilibrium and biological processes are not adequately
considered, and the speciation of selenium in oxidized natural waters is not accurately predicted. Selenate
is usually the predominate form in solution; however, selenite and organoselenium can both exist at
concentrations higher than predicted (Faust 1981; Luoma et al. 1997). Bioaccumulation by
microorganisms, bioproduction and release of organoselenium, and mineralization of particulate selenium
forms contribute to the disequilibrium.
4.0 Physical Distribution of Species in Surface Water	
The physical distribution of various selenium species in surface waters is regulated by:
	sorption to or incorporation in suspended particulate matter (SPM), and
	complexation with inorganic and/or organic colloidal material, such as (FeO OH)n and humic
substances (dissolved organic matter, DOM).
Both sorption to SPM and complexation with colloidal matter reduces the bioavailability of the
selenium species. The average fraction of selenium associated with the suspended particulate phase
(0.45|im filtration) as determined from eleven different studies of various surface waters was found to be
16% (0-39% range) of the total selenium, i.e., an average operationally defined dissolved selenium level
of 84% (Table A-l). In the James River, VA, the dissolved inorganic and organic selenium was found to
be 77% and 70% associated with colloidal matter, respectively (Takayangi and Wong 1984). A study of
lake ecosystems in Finland (Wang et al. 1995) found that 52% of the dissolved selenium was associated
with humic substances, and in a similar speciation study of Finnish stream waters, Lahermo et al. (1998)
determined that 36% of the selenium was complexed with humic matter. Hence, in various waterbodies
physical distribution as well as chemical speciation of selenium must be considered in relationship to
bioavailability and aquatic toxicity.
Until recently, the organic selenium fraction has been routinely measured as the difference
between total dissolved selenium and the sum of selenite and selenate. Unfortunately, the calculation of
this important selenium fraction in water as the difference between the total and measurable inorganic
fractions has not permitted this fraction to be fully characterized. New techniques are currently being
A-5

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developed which should help the specific identification and characterization of the nature of the organic
selenides present in aquatic systems. This work is particularly important because portions of the organic
selenium fraction (e.g., selenomethionine) of total dissolved selenium in water have been shown to be
much more bioavailable than the other forms of selenium, and therefore this work is also important for
understanding the manifestation of selenium ecotoxic effects.
Table A-l. Suspended particulate and dissolved selenium as a function of total selenium in
freshwater and marine aquatic ecosystems.			


Particulate So
Traction
Reference
Wsilerhodv
(% of Total)
dissolved, I'd
Cutter 1989
Carquinez, CA
20-40
0.6-0.8
Cutter 1986
Hyco Reservoir, NC
0
1
Tanizaki et al. 1992
Japanese Rivers
16
0.84
Luomaetal. 1992
San Francisco Bay, CA
22-31
0.69-0.78
Cumbie and VanHorn, 1978
Belews Lake, NC
8
0.92
GLEC 1997
Unnamed Stream, Albright, WV
4
0.96
Wang et al. 1995
Finnish Lakes
10
0.9
Lahermo et al. 1998
Finnish Streams
8
0.92
Hamilton et al. 200la,b
Adobe Creek, Fruita, CO
18
0.82
Hamilton et al. 200la,b
North Pond, Fruita, CO
0
1
Hamilton et al. 200la,b
Fish Ponds, Fruita, CO
7
0.93
Nakamoto and Hassler 1992
Merced River, CA
0
1
Nakamoto and Hassler 1992
Salt Slough, CA
4
0.96
Welsh and Maughan 1994
Cibola Lake, CA
39
0.62
Welsh and Maughan 1994
Hart Mine Marsh, Blythe, CA
6
0.94
Welsh and Maughan 1994
Colorado River, Blythe, CA
11
0.89
Welsh and Maughan 1994
Palo Verda Oxbow Lake, CA
33
0.67
Welsh and Maughan 1994
Palo Verda Outfall Drain, CA
0
1
Welsh and Maughan 1994
Pretty Water Lake, CA
21
0.79
A-6

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APPENDIX B: CONVERSIONS
B-l

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1.0 Conversion of Wet to Dry Tissue Weight
1.1 Methodology
Conversion factors (CF) derived from selenium measurements were calculated using
concentrations expressed as dry weights (jj.g/g dry weight). The majority of tissue and whole-body
selenium concentrations were reported as dry weights. Measurements reported as wet weight were
converted to equivalent dry weights using available percent moisture data for the relevant species and
tissue type.
Species-specific percent moisture data for muscle tissue were available for bluegill (Gillespie and
Baumann 1986; Nakamoto and Hassler 1992), rainbow trout (Seiler and Skorupa 2001), and for a
composite average of nine fish species (May et al. 2000). Species specific percent moisture data for
ovaries were available for bluegill (Gillespie and Baumann 1986; Nakamoto and Hassler 1992), fathead
minnow (GEI Associates 2008; Rickwood et al. 2008), and rainbow trout (Seiler and Skorupa 2001).
Species-specific % moisture data for whole-body tissues were available for bluegill (USGS NCBP).
Measurements reported as wet weight were converted to equivalent dry weights using available
percent moisture data for the relevant species and tissue type. If percent moisture data were unavailable
for a fish species, percent moisture data for a similar species (i.e., same genus or, if unavailable, same
family) were used. Table B-la lists percent moisture by tissue type, species, data source, and the target
species and study for which the % moisture data were used to convert from wet to dry weight. Table B-lb
is a list of 38 freshwater fish species and their percent solids and moisture. Although these data were not
needed for wet to dry weight conversion in any of the studies in this document, they are provided here as
a potential resource.
B-2

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% Moisture 1).
Species
Kii Source
Sliiily
% Moist
Wliolc-
hoilv
lire hv T
Muscle
ssue
()\sirv
C'omersion Appli
Species
eil to
Sliiily
Used in derivation of FCV
Rainbow trout
Seiler & Skorupa 2001


61.20
Rainbow trout
Holm et al. 2005
Rainbow trout
Seiler & Skorupa 2001


61.20
Brook trout
Holm et al. 2005
Fathead
minnow
Average of GEI Assoc. 2008;
Rickwood et al. 2008


75.30
Fathead minnow
Schultz and
Hermanutz 1990
Bluegill
Average of Gillespie &
Baumann 1986 andNakamoto
& Hassler 1992


76.00
Bluegill
Hermanutz et al.
1996
Avg of 9 spp
May et al. 2000

78.4

Striped bass
Coughlan and Velte
1989
Used in conversion of FCV in egg/ovary to whole-body Se concentrations
Bluegill
USGS NCBP
74.80


Bluegill
Hermanutz et al.
1996
Bluegill
May et al. 2000

80.09

Bluegill
Hermanutz et al.
1996
Bluegill
Average of Gillespie &
Baumann 1986 andNakamoto
& Hassler 1992


76.00
Bluegill
Hermanutz et al.
1996
Rainbow trout
May et al. 2000

77.54

Brook Trout
Holm et al. 2005
Rainbow trout
Seiler & Skorupa 2001


61.20
Brook Trout
Holm et al. 2005
Rainbow trout
May et al. 2000

77.54

Rainbow Trout
Holm et al. 2005
Rainbow trout
Seiler & Skorupa 2001


61.20
Rainbow Trout
Holm et al. 2005
Rainbow trout
May et al. 2000

77.54

Rainbow Trout
Casey & Siwik 2000
Rainbow trout
Seiler & Skorupa 2001


61.20
Rainbow Trout
Casey & Siwik 2000
B-3

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Table B-lb. Percent solids and moisture for whole body fish tissues by species.
Data provided by GEI Consultants (GEI2014).	
Species
A\era}c % solids
C'oiinl
Min
Max
A\ % moisture
Black bullhead
23.18
6
18.4
27
76.82
Blacknose dace
26.25
44
21.3
31.2
73.75
Bluntnose minnow
25.2
3
23.8
25.9
74.8
Brook stickleback
24.18
57
19.3
27.8
75.82
Carp
21.8
6
21.1
22.8
78.2
Central Stoneroller
25.38
174
17.2
33.7
74.62
Common carp
24.54
62
17.4
43
75.64
Creek chub
23.29
306
16.5
29.3
76.71
Fantail darter
27.71
15
19.5
72.3
72.29
Fathead minnow
23.36
298
15.3
100
76.64
Green sunfish
23.87
150
7.9
29
76.13
Greenside darter
25.55
11
21.7
27
74.45
Johnny darter
28.3
1
--
--
71.7
Largemouth bass
24.26
64
20.6
28.8
75.74
Log perch
23.05
2
22.3
23.8
76.95
Longnose dace
26.75
17
23.4
31.3
73.25
Mimic shiner
24.9
2
24
25.8
75.1
Mosquitofish
23.96
8
22.5
24
76.04
Northern hogsucker
23.93
113
17
39
76.07
Plains killifish
24.5
9
23.3
26.1
75.5
Rainbow darter
27.17
85
12
33.3
72.83
Red shiner
26.93
46
20.9
34.8
73.07
Redside shiner
24.44
8
21.8
26.9
75.56
River chub
24.8
4
22.9
27.3
75.2
River redhorse
20.8
1
--
--
79.2
Rock bass
25.05
24
21.2
29.3
74.95
Rosyface shiner
30.25
2
27.6
32.9
69.75
Rosyside shiner
24.54
5
23.1
25.7
75.46
Sand shiner
26.03
83
20.7
30.7
73.97
Sauger
23
1
--
--
77
Silver shiner
23.4
7
22.3
24.6
76.6
Smallmouth bass
25.78
12
22.7
28.1
74.22
Speckled dace
26.04
35
21
31.2
73.96
Striped shiner
22.9
64
18.2
28.8
77.1
Sunfish
23.2
1
-
-
76.8
Variegated darter
27.45
13
21.7
30.3
72.55
White sucker
22.63
246
16.5
28.4
77.37
Yellow perch
26.02
5
24
28.4
73.98
Grand total
24.85
1990


75.15
B-4

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2.0 Derivation of Tissue Conversion Factors
2.1 Methodology
EPA used a mechanistic bioaccumulation modeling approach to derive a mathematical
relationship between the concentration of selenium in water to the concentration of selenium in the eggs
and ovaries of fish. This approach characterizes selenium bioaccumulation as a series of steps
representing the phase transformation of selenium from dissolved to particulate form, and then the trophic
transfer of selenium through aquatic food webs to invertebrates and fish. The final step in this process is
the transfer of selenium into eggs and ovary tissue.
Equation 1 quantitatively models the transfer of selenium through each environmental
compartment as a series of site-specific and species-specific parameters. The parameter CF in Equation 1
represents the species-specific proportion of selenium in egg or ovary tissue relative to the average
concentration of selenium in all body tissues and is given as:
CF = C.g(Equation 1)
^whole-body
Where:
CF	= Whole-body to egg-ovary conversion factor (dimensionless ratio).
Cegg-ovary = Selenium concentration in the eggs or ovaries of fish (jj.g/g dw)
C whoie-body = Selenium concentration in the whole body of fish (|ag/g dw).
EPA derived species-specific conversion factor (CF) values using the same methods that were
used to derive species-specific TTF values from field data. To derive whole-body to egg-ovary CF values,
the EPA defined matched pairs of selenium measurements from the whole-body and from the eggs or
ovaries measured from the same individual fish or from matched composite samples. Egg-ovary
concentration was defined as a measurement from either the eggs or the ovaries. If multiple measurements
from both eggs and ovaries of the same individual or matched composite sample were available, the
average value was used. EPA first confirmed a statistical relationship between egg-ovary and whole body
selenium for each species using ordinary least squares (OLS) linear regression. If the regression resulted
in a statistically significant (P<0.05) positive slope, EPA calculated the ratio of the egg-ovary to whole
body selenium concentration for each matched pair of measurements and used the median as the CF value
for that species.
EPA derived CF values from selenium measurements in units of |_ig/g dry weight. The majority of
tissue and whole body selenium concentrations were reported as dry weights. Measurements reported as
B-5

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wet weight were converted to equivalent dry weights using available percent moisture data for the
relevant species and tissue type. If percent moisture data were unavailable for a fish species, percent
moisture data for a similar species (i.e., same genus or, if unavailable, same family) were used. A listing
of percent moisture concentrations by species and target tissue are provided in Table B-la.
For those species without sufficient data to directly calculate an egg-ovary to whole body CF, but
which had sufficient data to calculate a conversion factor for either egg-ovary to muscle or whole body to
muscle, EPA followed a two stage approach based on taxonomic similarity, similar to that described
above. If a fish species had species specific egg-ovary to muscle conversion factor, but no whole body
data with which to calculate an egg to whole body CF, then available data would be used to estimate a
muscle to whole body conversion factor for that species based on taxonomic relatedness. The estimated
muscle to whole body factor would be multiplied by the directly measured egg-ovary to muscle factor to
estimate an egg-ovary to whole body CF for that species. For example, rainbow trout has a species
specific egg-ovary to muscle conversion factor of 1.92, but does not have a species specific egg-ovary to
whole body CF. Using the taxonomic approach described above, the most closely related taxa to rainbow
trout with muscle to whole body conversion factors are in the class Actinopterygii. The median
conversion factor for the 8 species within that class is 1.27. The final egg-ovary to whole body CF for
rainbow trout is 2.44 (Table B-6), or 1.92 x 1.27.
The EPA developed species-specific egg-ovary to muscle and muscle to whole-body correction
factors following the procedure described for whole-body to egg-ovary conversion factors. The EPA
obtained matched pairs of selenium measurements in the whole-body and muscle filets and matched pairs
of selenium measurements in muscle filets and whole-body from published scientific literature. EPA first
confirmed a statistical relationship between the two tissue types for each species using OLS linear
regression. If the regression resulted in a significant fit with a positive slope, the EPA calculated the ratio
of each matched pair of measurements and then calculated the median ratio.
B-6

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2.2 CF values calculated directly from whole-body and egg-ovary selenium measurements
n
^whole-body
r
^-egg
c
^ ovary
r
v-- poo-nvarv
Ratio
Selenium concentration in all tissues (jj.g/g dw)
Selenium concentration in eggs (jj.g/g dw)
Selenium concentration in ovary tissue (jj.g/g dw)
Average selenium concentration in eggs and ovaries
C
egg-ovary
c +c ^
egg	ovary
C
whole-body
Black bullhead (Ameiurus melas)
Study
Cwhole-bodv
r
e2S
r
^ovarv
r
*^egg-ovarv
Ratio
Osmundson et al. 2007
5.30

64.30
64.30
12.13
Osmundson et al. 2007
4.80
-
35.40
35.40
7.38
Osmundson et al. 2007
5.50
-
52.80
52.80
9.60
Osmundson et al. 2007
4.90
-
56.00
56.00
11.43
Osmundson et al. 2007
9.60
-
42.80
42.80
4.46
Osmundson et al. 2007
7.60
-
38.70
38.70
5.09
Osmundson et al. 2007
7.30
-
37.30
37.30
5.11
Osmundson et al. 2007
6.60
-
34.30
34.30
5.20
Osmundson et al. 2007
8.60
-
26.40
26.40
3.07
Osmundson et al. 2007
2.00
-
56.70
56.70
28.35
Osmundson et al. 2007
5.30
-
64.30
64.30
12.13
60 n
40 -
20 -
20
c
40
whole-body
Median ratio: 6.29
R2:
0.37
F:
4.67
df:
8
P:
0.046
Not used because negative slope.
60
B-7

-------
Bluegill (Lepomis macrochirus)
Study
Cwhole-bodv
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Coyle et al. 1993
0.90
1.90
2.10
2.00
2.22
Coyle et al. 1993
2.90
7.30
8.30
7.80
2.69
Coyle et al. 1993
4.90
13.00
12.50
12.75
2.60
Coyle et al. 1993
7.20
22.80
25.00
23.90
3.32
Coyle et al. 1993
16.00
41.30
41.00
41.15
2.57
Doroshov et al. 1992
1.60
2.80
-
2.80
1.75
Doroshov et al. 1992
5.50
8.30
-
8.30
1.51
Doroshov et al. 1992
9.30
19.50
-
19.50
2.10
Doroshov et al. 1992
19.30
38.40
-
38.40
1.99
Hermanutz et al. 1996
1.50
-
0.30
0.30
0.20
Hermanutz et al. 1996
18.10
-
16.70
16.70
0.92
Hermanutz et al. 1996
1.90
-
4.40
4.40
2.32
Hermanutz et al. 1996
2.80
-
8.40
8.40
3.00
Hermanutz et al. 1996
12.30
-
29.00
29.00
2.36
Hermanutz et al. 1996
9.40
-
24.50
24.50
2.61
Hermanutz et al. 1996
1.50
-
3.20
3.20
2.13
Hermanutz et al. 1996
4.90
-
10.30
10.30
2.10
Hermanutz et al. 1996
21.00
-
42.10
42.10
2.00
Hermanutz et al. 1996
24.30
-
55.00
55.00
2.26
Hermanutz et al. 1996
5.00
-
7.00
7.00
1.40
Hermanutz et al. 1996
9.50
-
26.00
26.00
2.74
Hermanutz et al. 1996
6.60
-
14.90
14.90
2.26
Hermanutz et al. 1996
1.80
-
4.40
4.40
2.44
Hermanutz et al. 1996
4.20
-
7.90
7.90
1.88
Hermanutz et al. 1996
10.30
-
16.30
16.30
1.58
Hermanutz et al. 1996
13.80
-
15.90
15.90
1.15
Osmundson et al. 2007
8.80
-
9.70
9.70
1.10
60 n
'egg-ovary
Median ratio: 2.13
R
F
df
P
0.82
110.9
25
<0.001
-whole-body
B-8

-------
Bluehead sucker (Catostomus discobolus)
Study
Cwhole-bodv
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
1.30

2.40
2.40
1.85
Osmundson et al. 2007
2.00
-
4.20
4.20
2.10
Osmundson et al. 2007
2.10
-
3.70
3.70
1.76
Osmundson et al. 2007
2.20
-
4.00
4.00
1.82
Osmundson et al. 2007
2.40
-
4.10
4.10
1.71
Osmundson et al. 2007
3.90
-
7.10
7.10
1.82
Osmundson et al. 2007
5.60
-
8.10
8.10
1.45
Median ratio
egg-ovary
whole-body
Brown trout (Salmo trutta)
Study

Cwhole-bodv
r
v^egg
c c
^ovarv ^ess-ovarv
Ratio
Formation 201
Saratoga fish hatchery
3.60
0.80
0.80
0.22
Formation 201
Saratoga fish hatchery
4.10
0.90
0.90
0.22
Formation 201
Saratoga fish hatchery
3.70
0.80
0.80
0.22
Formation 201
Saratoga fish hatchery
4.30
0.90
0.90
0.21
Formation 201
Saratoga fish hatchery
3.00
1.20
1.20
0.40
Formation 201
Saratoga fish hatchery
3.10
1.20
1.20
0.39
Formation 201
Saratoga fish hatchery
2.70
1.00
1.00
0.37
Formation 201
Saratoga fish hatchery
2.50
1.00
1.00
0.40
Formation 201
Saratoga fish hatchery
8.90
12.80
12.80
1.44
Formation 201

13.80
40.30
40.30
2.92
Formation 201

17.90
36.00
36.00
2.01
Formation 201

13.60
26.80
26.80
1.97
Formation 201

17.20
26.90
26.90
1.56
Formation 201

6.70
18.60
18.60
2.78
Formation 201

9.60
17.70
17.70
1.84
Formation 201

22.60
38.80
38.80
1.72
Formation 201

7.20
13.20
13.20
1.83
Formation 201

9.20
13.40
13.40
1.46
B-9

-------
Brown trout (Salmo trutta)
Formation 2011
13.20
Formation 2011
8.60
Formation 2011
11.30
Formation 2011
20.00
Formation 2011
8.40
Formation 2011
5.60
Formation 2011
6.70
Formation 2011
5.90
Formation 2011
6.00
Formation 2011
7.00
Formation 2011
5.60
Formation 2011
4.70
Formation 2011
7.20
Formation 2011
9.20
Formation 2011
5.50
Formation 2011
8.50
Osmundson et al. 2007
4.60
Osmundson et al. 2007
4.30
Osmundson et al. 2007
5.00
Osmundson et al. 2007
5.50
C whole-body
20.50
-
20.50
1.55
12.50
-
12.50
1.45
11.20
-
11.20
0.99
28.10
-
28.10
1.41
12.80
-
12.80
1.52
8.40
-
8.40
1.50
8.50
-
8.50
1.27
8.40
-
8.40
1.42
9.10
-
9.10
1.52
7.50
-
7.50
1.07
6.60
-
6.60
1.18
6.90
-
6.90
1.47
6.20
-
6.20
0.86
14.00
-
14.00
1.52
6.90
-
6.90
1.25
9.50
-
9.50
1.12
-
1.20
1.20
0.26
-
37.80
37.80
8.79
-
35.60
35.60
7.12
-
32.50
32.50
5.91
Median ratio:	1.45
R2: 0.47
F:	31.3
df:	36
P:	<0.001
B-10

-------
Channel catfish (Ictalurus punctatus)
Study
Cwhole-bodv
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
3.40

29.50
29.50
8.68
Osmundson et al. 2007
3.30
-
21.10
21.10
6.39
Osmundson et al. 2007
2.60
-
13.70
13.70
5.27
Osmundson et al. 2007
4.00
-
30.30
30.30
7.58
30 -
20 
10 -
Median ratio: 6.98
R2:
0.82
F:
9.1
df:
2
P:
0.099
Not used because P > 0.05.
10
20
-whole-body
30
Common carp (Cyprinus carpio)
Study
Cwhole-bodv
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
6.30

12.10
12.10
1.92
Osmundson et al. 2007
4.80
-
9.40
9.40
1.96
Osmundson et al. 2007
11.70
-
16.30
16.30
1.39
Osmundson et al. 2007
23.10
-
27.30
27.30
1.18
Osmundson et al. 2007
4.10
-
9.90
9.90
2.41
30
20
'egg-ovary
10
10
20
-whole-body
Median ratio: 1.92
R
F
df
P
30
0.96
584.8
3
<0.001
B-ll

-------
Creek chub (Semotilus atromaculatus)
Study
Cwhole-bodv
r
v^egg
C C
^ovarv ^ egg-ova rv
Ratio
GEI2014
2.89
6.86
6.86
2.37
GEI2014
4
9.94
9.94
2.49
GEI 2014
4.14
8.1
8.1
1.96
GEI 2014
4.46
8.98
8.98
2.01
GEI 2014
5.57
18.63
18.63
3.34
GEI 2014
6.23
22.35
22.35
3.59
GEI 2014
24.26
39.07
39.07
1.61
GEI 2014
20.49
12.38
12.38
0.60
GEI 2014
16.33
19.59
19.59
1.20
GEI 2014
14.03
23.78
23.78
1.69
GEI 2014
5.71
23.21
23.21
4.06
GEI 2014
8.17
16.03
16.03
1.96
i'gg ovary
45
40
35
30
25
20
15
10
5
0
10 15
20 25
r
^whnlp-hnih
30 35 40 45
Median ratio: 1.99
R
F
df
P
0.82
7.09
10
0.012
Desert pupfish (Cyprinoclon macularius)
Study
r
^whole-bodv
r
egg
C C
^ ovary ^ egg-ova rv
Ratio
Besser et al. 2012
0.75
1
l
1.33
Besser et al. 2012
2.5
3
3
1.20
Besser et al. 2012
3.4
4.4
4.4
1.29
Besser et al. 2012
6.7
8
8
1.19
Besser et al. 2012
12
13
13
1.08
Besser et al. 2012
24
27
27
1.13
B-12

-------
Desert pupfish (Cyprinodon macularius)
30
25
20
'15
10
o

10 15 20
r
^--whnlp-hfifiv
25
30
Median ratio: 1.20
R
F
df
P
1.00
194.3
4
<0.001
Cutthroat trout (Oncorhynchus clarkii)
Study
Cwhole-bodv
r
egg
C C
^ovarv ^egg-ovary
Ratio
Hardy 2005
0.70
1.00
1.00
1.43
Hardy 2005
2.60
3.80
3.80
1.46
Hardy 2005
2.80
5.50
5.50
1.96
Hardy 2005
6.40
18.00
18.00
2.81
Hardy 2005
1.20
1.60
1.60
1.33
Hardy 2005
4.60
7.80
7.80
1.70
Hardy 2005
5.90
6.60
6.60
1.12
Hardy 2005
9.10
5.10
5.10
0.56
Hardy 2005
11.40
5.20
5.20
0.46
Hardy 2005
5.60
16.00
16.00
2.86
Formation 2012
2.56
3.43
3.43
1.34
Formation 2012
16.3
17.6
17.6
1.08
Formation 2012
20.7
27.9
27.9
1.35
Formation 2012
19.4
29.7
29.7
1.53
Formation 2012
17
22.3
22.3
1.31
Formation 2012
16.7
14.6
14.6
0.87
Formation 2012
25.7
47.6
47.6
1.85
Formation 2012
8.17
22
22
2.69
Formation 2012
9.07
15.4
15.4
1.70
Formation 2012
8.63
11.4
11.4
1.32
Formation 2012
16.6
12.7
12.7
0.77
Formation 2012
19.4
40.1
40.1
2.07
Formation 2012
21
30
30
1.43
Formation 2012
18.6
35.6
35.6
1.91
Formation 2012
22.5
30.5
30.5
1.36
Formation 2012 Henry Lake fish hatchery
0.4
1.65
1.65
4.13
B-13

-------
Cutthroat trout (Oncorhynchus clarkii)
Formation 2012 Henry Lake fish hatchery
0.45
2.03
2.03
4.51
Formation 2012 Henry Lake fish hatchery
0.44
2.48
2.48
5.64
Formation 2012 Henry Lake fish hatchery
0.36
1.36
1.36
3.78
Formation 2012 Henry Lake fish hatchery
0.5
2.33
2.33
4.66
Formation 2012 Henry Lake fish hatchery
0.36
0.83
0.83
2.31
Formation 2012 Henry Lake fish hatchery
0.44
2.26
2.26
5.14
Formation 2012 Henry Lake fish hatchery
0.28
1.87
1.87
6.68
Formation 2012 Henry Lake fish hatchery
0.44
1.98
1.98
4.50
Formation 2012 Henry Lake fish hatchery
0.43
1.34
1.34
3.12
Formation 2012 Henry Lake fish hatchery
0.31
3.23
3.23
10.42
Formation 2012 Henry Lake fish hatchery
0.23
1.58
1.58
6.87
Formation 2012 Henry Lake fish hatchery
0.72
1.93
1.93
2.68
Formation 2012 Henry Lake fish hatchery
0.73
1.79
1.79
2.45
Formation 2012 Henry Lake fish hatchery
0.91
2.06
2.06
2.26
Formation 2012 Henry Lake fish hatchery
0.85
1.74
1.74
2.05
-egg-c
Median ratio:
1.96
R2:
0.83
F:
194.3
df:
39
P:
<0.001
-whole-body
Fathead minnow (Pimephales promelas)
Study
Cwhole-bodv
r
v^egg
C C
^ovarv ^ess-ovarv
Ratio
GEI2014
2.04
3.81
3.81
1.87
GEI2014
1.39
2.23
2.23
1.60
GEI 2014
1.85
3.31
3.31
1.79
GEI 2014
1.32
3.43
3.43
2.60
GEI 2014
1.55
3.08
3.08
1.99
GEI 2014
37.13
50.06
50.06
1.35
GEI 2014
29.54
37.77
37.77
1.28
GEI 2014
33.32
40.82
40.82
1.23
GEI 2014
28.26
32.23
32.23
1.14
GEI 2014
30.74
46.21
46.21
1.50
B-14

-------
Fathead minnow (Pimephales promelas)
GEI2014
53.17
60.84
60.84
1.14
GEI2014
48.52
39.28
39.28
0.81
GEI 2014
53.81
44.28
44.28
0.82
GEI 2014
53.2
46.21
46.21
0.87
GEI 2014
54.01
43.51
43.51
0.81
GEI 2014
12.93
23.18
23.18
1.79
GEI 2014
8.19
14.67
14.67
1.79
GEI 2014
14.25
32.04
32.04
2.25
GEI 2014
8.65
19.95
19.95
2.31
GEI 2014
16.33
38.51
38.51
2.36
GEI 2014
7.69
7.39
7.39
0.96
GEI 2014
19.05
29.69
29.69
1.56
GEI 2014
8.78
9.55
9.55
1.09
GEI 2014
14.68
36.58
36.58
2.49
GEI 2014
9.02
13.63
13.63
1.51
GEI 2014
46.17
61.99
61.99
1.34
GEI 2014
41.97
60.07
60.07
1.43
GEI 2014
34.33
42.74
42.74
1.24
GEI 2014
33.4
38.89
38.89
1.16
GEI 2014
42.53
71.24
71.24
1.68
GEI 2014
74.56
85.87
85.87
1.15
GEI 2014
67.94
65.85
65.85
0.97
GEI 2014
70.85
58.91
58.91
0.83
GEI 2014
43.93
49.67
49.67
1.13
GEI 2014
66.57
67.39
67.39
1.01
GEI 2014
20.21
58.91
58.91
2.91
GEI 2014
13.08
65.85
65.85
5.03
GEI 2014
23.02
31.38
31.38
1.36
GEI 2014
11.55
25.72
25.72
2.23
GEI 2014
32.8
48.52
48.52
1.48
GEI 2014
27.17
48.9
48.9
1.80
GEI 2014
28.54
38.04
38.04
1.33
GEI 2014
37.2
73.16
73.16
1.97
GEI 2014
32.79
44.28
44.28
1.35
GEI 2014
46.17
61.99
61.99
1.87
B-15

-------
Fathead minnow (Pimephales promelas)
100
60
40
20
%
6>

-------
Green sunfish (Lepomis cyanellus)
Study
r
^whole-bodv
r
ess
c
v o\ ;tr\
r
^ess-ovarv
Ratio
Osmundson et al. 2007
22.80

27.40
27.40
1.20
Osmundson et al. 2007
8.80
-
10.20
10.20
1.16
Osmundson et al. 2007
15.40
-
21.80
21.80
1.42
Osmundson et al. 2007
4.80
-
7.00
7.00
1.46
Osmundson et al. 2007
5.70
-
8.90
8.90
1.56
Osmundson et al. 2007
4.40
-
6.40
6.40
1.45
Osmundson et al. 2007
3.80
-
6.40
6.40
1.68
Osmundson et al. 2007
11.90
-
18.10
18.10
1.52
Osmundson et al. 2007
6.40
-
12.30
12.30
1.92
Osmundson et al. 2007
9.50
-
13.80
13.80
1.45
Osmundson et al. 2007
9.10
-
15.20
15.20
1.67
Osmundson et al. 2007
6.20
-
10.80
10.80
1.74
Osmundson et al. 2007
7.00
-
11.70
11.70
1.67
Osmundson et al. 2007
7.70
-
12.60
12.60
1.64
Osmundson et al. 2007
6.20
-
10.00
10.00
1.61
Osmundson et al. 2007
10.20
-
13.90
13.90
1.36
Osmundson et al. 2007
9.70
-
15.20
15.20
1.57
Osmundson et al. 2007
9.90
-
14.70
14.70
1.48
Osmundson et al. 2007
7.20
-
8.80
8.80
1.22
Osmundson et al. 2007
9.00
-
12.90
12.90
1.43
Osmundson et al. 2007
9.70
-
13.10
13.10
1.35
Osmundson et al. 2007
8.90
-
11.50
11.50
1.29
Osmundson et al. 2007
9.80
-
13.20
13.20
1.35
Osmundson et al. 2007
9.90
-
11.60
11.60
1.17
Osmundson et al. 2007
10.30
-
7.50
7.50
0.73
Osmundson et al. 2007
5.30
-
8.10
8.10
1.53
Osmundson et al. 2007
10.10
-
13.20
13.20
1.31
Osmundson et al. 2007
11.80
-
14.00
14.00
1.19
Osmundson et al. 2007
3.30
-
5.20
5.20
1.58
Osmundson et al. 2007
4.00
-
5.80
5.80
1.45
Osmundson et al. 2007
4.30
-
4.10
4.10
0.95
Osmundson et al. 2007
3.70
-
4.90
4.90
1.32
Osmundson et al. 2007
6.20
-
9.50
9.50
1.53
Osmundson et al. 2007
3.50
-
4.80
4.80
1.37
Osmundson et al. 2007
4.40
-
5.60
5.60
1.27
Osmundson et al. 2007
5.60
-
10.10
10.10
1.80
Osmundson et al. 2007
4.90
-
7.50
7.50
1.53
Osmundson et al. 2007
4.40
-
5.90
5.90
1.34
B-17

-------
Green sunfish (Lepomis cyanellus)
30
20
'egg-ovary
10
Median ratio: 1.45
R
F
df
P
0.87
240.0
36
<0.001
10
20
30
"whole-body
Roundtail chub (Gila robusta)
Study
Cwhole-bodv
r
V'egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
4.10

7.90
7.90
1.93
Osmundson et al. 2007
5.30
-
10.80
10.80
2.04
Osmundson et al. 2007
6.40
-
15.20
15.20
2.38
Osmundson et al. 2007
6.80
-
14.10
14.10
2.07
Osmundson et al. 2007
5.50
-
10.60
10.60
1.93
Osmundson et al. 2007
6.60
-
18.00
18.00
2.73
Osmundson et al. 2007
8.40
-
17.80
17.80
2.12
20
10 
'egg-ovary
O fo
Median ratio: 2.07
R2:
0.80
F:
20.4
df:
5
P:
0.004
10
20
"whole-body
B-18

-------
Smallmouth bass (Micropterus dolomieu)
Study
Cwhole-bodv
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
4.20

6.00
6.00
1.43
Osmundson et al. 2007
5.50
-
8.00
8.00
1.45
Osmundson et al. 2007
5.40
-
6.50
6.50
1.20
Osmundson et al. 2007
7.80
-
11.00
11.00
1.41
Osmundson et al. 2007
5.10
-
7.10
7.10
1.39
Osmundson et al. 2007
4.90
-
8.80
8.80
1.80
Median ratio
egg-ovary
whole-body
White sucker (Catostomus commersonii)
Study
Cwhole-bodv
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
3.80

6.20
6.20
1.63
Osmundson et al. 2007
4.20
-
6.20
6.20
1.48
Osmundson et al. 2007
3.30
-
5.20
5.20
1.58
Osmundson et al. 2007
4.50
-
6.50
6.50
1.44
Osmundson et al. 2007
6.30
-
7.70
7.70
1.22
Osmundson et al. 2007
6.80
-
5.80
5.80
0.85
Osmundson et al. 2007
11.00
-
10.90
10.90
0.99
Osmundson et al. 2007
12.70
-
11.20
11.20
0.88
Osmundson et al. 2007
5.70
-
9.40
9.40
1.65
Osmundson et al. 2007
3.90
-
5.40
5.40
1.38
Osmundson et al. 2007
3.80
-
5.10
5.10
1.34
Osmundson et al. 2007
9.90
-
10.40
10.40
1.05
Osmundson et al. 2007
5.30
-
10.40
10.40
1.96
Osmundson et al. 2007
10.70
-
11.00
11.00
1.03
Osmundson et al. 2007
5.90
-
11.70
11.70
1.98
Osmundson et al. 2007
7.00
-
11.60
11.60
1.66
Osmundson et al. 2007
6.40
-
9.40
9.40
1.47
Osmundson et al. 2007
6.30
-
10.20
10.20
1.62
Osmundson et al. 2007
5.30
-
7.30
7.30
1.38
Osmundson et al. 2007
6.20
-
8.90
8.90
1.44
B-19

-------
White sucker (Catostomus commersonii)
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
GEI2014
GEI2014
40
30
10
0
0
10
30
40
f
^whole-body
5.60
8.80
8.70
11.40
10.70
8.40
7.00
7.50
10.30
6.70
2.10
1.80
3.20
2.30
3.10
3.00
2.80
2.50
3.40
2.80
26.9
22.9
10.50
10.50
1.88
10.20
10.20
1.16
8.10
8.10
0.93
9.50
9.50
0.83
10.70
10.70
1.00
8.30
8.30
0.99
12.00
12.00
1.71
6.10
6.10
0.81
6.10
6.10
0.59
11.30
11.30
1.69
2.60
2.60
1.24
3.60
3.60
2.00
4.40
4.40
1.38
4.40
4.40
1.91
4.80
4.80
1.55
4.30
4.30
1.43
4.10
4.10
1.46
3.80
3.80
1.52
3.60
3.60
1.06
3.80
3.80
1.36
32.7
32.7
1.22
23.3
23.3
1.02
Median ratio:	1.38
R2:	0.83
F:	200.4
df:	40
P:	<0.001
B-20

-------
Table B-2. Summary of egg-ovary to whole body conversion factors (CF) from matched pairs of
whole-body and egg-ovary measurements.	
(0111111011 name
Scientific name
Median ralio {( '/')
Bluegill
Lepomis macrochirus
2.13
Bluehead sucker
Catostomus discobolus
1.82
Brown trout
Salmo trutta
1.45
Common carp
Cyprinus carpio
1.92
Creek chub
Semotilus atromaculatus
1.99
Cutthroat trout
Oncorhynchus clarkii
1.96
Desert pupfish
Cyprinodon macularius
1.20
Fathead minnow
Pimephales promelas
1.40
Flannelmouth sucker
Catostomus latipinnis
1.41
Green sunfish
Lepomis cyanellus
1.45
Roundtail chub
Gila robusta
2.07
Smallmouth bass
Micropterus dolomieu
1.42
White sucker
Catostomus commersonii
1.38
B-21

-------
2.3 Muscle to egg-ovary conversion factors
Cmuscle = Selenium concentration in muscle tissue only ((.ig/g dw)
Cegg	= Selenium concentration in eggs ((.ig/g dw)
Covary = Selenium concentration in ovary tissue ((.ig/g dw)
fc^+c
= Average selenium concentration in eggs and ovaries
Ct
egg-ovary
egg
ovary
Ratio
C
egg-ovdx y
C
muscle
Black bullhead (Ameiurus melas)
Study
Cmuscle
r
egg
r
^ovarv
r
*^egg-ovarv
Ratio
Osmundson et al. 2007
3.40

64.30
64.30
18.91
Osmundson et al. 2007
3.90
-
35.40
35.40
9.08
Osmundson et al. 2007
4.30
-
52.80
52.80
12.28
Osmundson et al. 2007
4.70
-
56.00
56.00
11.91
Osmundson et al. 2007
5.70
-
42.80
42.80
7.51
Osmundson et al. 2007
7.40
-
38.70
38.70
5.23
Osmundson et al. 2007
7.50
-
37.30
37.30
4.97
Osmundson et al. 2007
7.80
-
34.30
34.30
4.40
Osmundson et al. 2007
7.80
-
26.40
26.40
3.38
Osmundson et al. 2007
9.20
-
56.70
56.70
6.16
60
f	4(1
^egg-ovary
4	6
^"muscle
Median ratio: 6.84
R2:
0.17
F:
1.65
df:
8
P:
0.250
10
Not used because P > 0.05 and negative
slope.
B-22

-------
Bluegill (Lepomis macrochirus)
Study

Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Brysonetal. 1984
84.0

49.0
49.0
0.58
Brysonetal. 1985a (pt. 1)
59.0
-
30.0
30.0
0.51
Brysonetal. 1985a (pt. 1)
2.7
-
2.2
2.2
0.81
Bryson et al. 1985a (pt. 2)
25.0
-
9.1
9.1
0.36
Doroshov et al.
1992
1.5
2.8
-
2.8
1.87
Doroshov et al.
1992
5.8
8.3
-
8.3
1.43
Doroshov et al.
1992
10.4
19.5
-
19.5
1.88
Doroshov et al.
1992
23.6
38.4
-
38.4
1.63
Hermanutz et al
1996
1.6
-
2.0
2.0
1.25
Hermanutz et al
1996
8.5
-
18.8
18.8
2.21
Hermanutz et al
1996
14
-
15.5
15.5
1.11
Hermanutz et al
1996
2.1
-
0.3
0.3
0.14
Hermanutz et al
1996
20.6
-
16.7
16.7
0.81
Hermanutz et al
1996
1.9
-
4.4
4.4
2.32
Hermanutz et al
1996
3.5
-
8.4
8.4
2.40
Hermanutz et al
1996
17.6
-
29.0
29.0
1.65
Hermanutz et al
1996
12.5
-
24.5
24.5
1.96
Hermanutz et al
1996
2.3
-
3.2
3.2
1.39
Hermanutz et al
1996
6.9
-
10.3
10.3
1.49
Hermanutz et al
1996
44.9
-
42.1
42.1
0.94
Hermanutz et al
1996
39.8
-
55.0
55.0
1.38
Hermanutz et al
1996
5.3
-
7.0
7.0
1.32
Hermanutz et al
1996
12.5
-
26.0
26.0
2.08
Hermanutz et al
1996
7.8
-
14.9
14.9
1.91
Hermanutz et al
1996
3.2
-
4.4
4.4
1.38
Hermanutz et al
1996
6.1
-
7.9
7.9
1.30
Hermanutz et al
1996
18.7
-
16.3
16.3
0.87
Hermanutz et al
1996
15.1
-
15.9
15.9
1.05
Osmundson et al. 2007
12.9
-
9.7
9.7
0.75
Median ratio
egg-ovary
muscle
B-23

-------
Bluehead sucker (Catostomus discobolus)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
1.5

2.4
2.4
1.60
Osmundson et al. 2007
2.3
-
4.2
4.2
1.83
Osmundson et al. 2007
2.5
-
3.7
3.7
1.48
Osmundson et al. 2007
2.7
-
4
4
1.48
Osmundson et al. 2007
3.1
-
4.1
4.1
1.32
Osmundson et al. 2007
5.2
-
7.1
7.1
1.37
Osmundson et al. 2007
8.6
-
8.1
8.1
0.94
Median ratio
egg-ovary
muscle
Brook trout (Salvelinus fontinalis)
Study
r
^muscle
r
egg
C C
^ovarv ^ess-ovarv
Ratio
Holm et al. 2005
2.80
1.50
1.50
0.54
Holm et al. 2005
1.40
2.50
2.50
1.79
Holm et al. 2005
2.20
3.40
3.40
1.55
Holm et al. 2005
2.00
4.70
4.70
2.35
Holm et al. 2005
2.20
2.90
2.90
1.32
Holm et al. 2005
5.00
5.60
5.60
1.12
Holm et al. 2005
9.70
9.90
9.90
1.02
Holm et al. 2005
10.50
15.40
15.40
1.47
Holm et al. 2005
11.20
12.80
12.80
1.14
Holm et al. 2005
11.40
14.80
14.80
1.30
Holm et al. 2005
12.30
12.20
12.20
0.99
Holm et al. 2005
15.90
12.40
12.40
0.78
Holm et al. 2005
16.50
13.20
13.20
0.80
Holm et al. 2005
19.60
15.50
15.50
0.79
Holm et al. 2005
20.40
15.30
15.30
0.75
Holm et al. 2005
23.40
25.40
25.40
1.09
Holm et al. 2005
34.70
32.50
32.50
0.94
B-24

-------
Brook trout (Salvelinus fontinalis)
Median ratio
egg-ovary
muscle
Brown trout (Salmo trutta)
Study
Cmuscle
r
egg
r
^ovarv
r
^egg-ovary
Ratio
Osmundson et al. 2007
3.2

1.2
1.2
0.38
Osmundson et al. 2007
3.6
-
37.8
37.8
10.50
Osmundson et al. 2007
4
-
35.6
35.6
8.90
Osmundson et al. 2007
6.3
-
32.5
32.5
5.16
C 20
egg-ovary
muscle
Median ratio: 7.03
R2:
0.17
F:
0.40
df:
2
P:
0.71
Not used because P > 0.05.
B-25

-------
Channel catfish (Ictaluris punctatus)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
3.4

29.5
29.5
8.68
Osmundson et al. 2007
3.6
-
21.1
21.1
5.86
Osmundson et al. 2007
3.7
-
13.7
13.7
3.70
Osmundson et al. 2007
5.3
-
30.3
30.3
5.72
X
ovary
muscle
Median ratio: 5.79
R2:
0.20
F:
0.49
df:
2
P:
0.67
Not used because P > 0.05.
Common carp (Cyprinus carpio)
Study
Cmuscle
r
egg
r
^ovarv
r
^egg-ovarv
Ratio
Garcia-Hernandez 2000
4.6

1.8
1.8
0.39
Osmundson et al. 2007
7.8
-
12.1
12.1
1.55
Osmundson et al. 2007
8.2
-
9.4
9.4
1.15
Osmundson et al. 2007
20
-
16.3
16.3
0.82
Osmundson et al. 2007
24.2
-
27.3
27.3
1.13
Osmundson et al. 2007
6.6
-
9.9
9.9
1.50
ovary
muscle
Median ratio: 1.14
R2:
0.84
F:
21.7
df:
4
P:
0.007
B-26

-------
Cutthroat trout (Oncorhynchus clarkii)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Golder 2005
6.80

28.20
28.20
4.15
Golder 2005
4.20
-
47.80
47.80
11.38
Golder 2005
3.00
-
22.00
22.00
7.33
Golder 2005
4.90
-
9.80
9.80
2.00
Golder 2005
4.50
-
8.20
8.20
1.82
Golder 2005
4.00
-
7.00
7.00
1.75
Golder 2005
5.00
-
10.00
10.00
2.00
Golder 2005
5.00
-
10.00
10.00
2.00
Golder 2005
5.00
-
8.00
8.00
1.60
Golder 2005
8.40
-
16.20
16.20
1.93
Golder 2005
8.30
-
18.30
18.30
2.20
Golder 2005
7.00
-
14.30
14.30
2.04
Golder 2005
6.60
-
14.30
14.30
2.17
Golder 2005
8.40
-
14.70
14.70
1.75
Golder 2005
9.80
-
16.40
16.40
1.67
Golder 2005
8.50
-
15.90
15.90
1.87
Golder 2005
16.00
-
20.00
20.00
1.25
Golder 2005
7.00
-
14.00
14.00
2.00
Golder 2005
8.00
-
19.00
19.00
2.38
Golder 2005
7.00
-
14.00
14.00
2.00
Golder 2005
7.00
-
14.00
14.00
2.00
Golder 2005
9.00
-
16.00
16.00
1.78
Golder 2005
7.00
-
13.00
13.00
1.86
Golder 2005
7.00
-
14.00
14.00
2.00
Golder 2005
8.00
-
14.00
14.00
1.75
Golder 2005
9.80
-
20.20
20.20
2.06
Golder 2005
7.00
-
22.00
22.00
3.14
Golder 2005
9.00
-
16.00
16.00
1.78
Golder 2005
7.00
-
12.00
12.00
1.71
Golder 2005
8.00
-
13.00
13.00
1.63
Golder 2005
10.00
-
14.00
14.00
1.40
Kennedy et al. 2000
41.30
75.40
66.80
71.10
1.72
Kennedy et al. 2000
15.30
58.40
31.60
45.00
2.94
Kennedy et al. 2000
14.10
30.60
31.40
31.00
2.20
Kennedy et al. 2000
12.50
20.20
18.50
19.35
1.55
Kennedy et al. 2000
13.70
19.40
19.50
19.45
1.42
Kennedy et al. 2000
14.30
16.20
16.20
16.20
1.13
Kennedy et al. 2000
9.50
16.10
19.30
17.70
1.86
Kennedy et al. 2000
9.40
14.40
22.00
18.20
1.94
Kennedy et al. 2000
8.70
13.20
17.00
15.10
1.74
Kennedy et al. 2000
9.50
12.60
13.60
13.10
1.38
Kennedy et al. 2000
10.20
12.30
14.50
13.40
1.31
B-27

-------
Cutthroat trout (Oncorhynchus clarkii)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Kennedy et al. 2000
10.70
10.50
20.60
15.55
1.45
Kennedy et al. 2000
6.60
9.90
21.50
15.70
2.38
Kennedy et al. 2000
9.70
9.10
13.20
11.15
1.15
Kennedy et al. 2000
10.90
8.50
13.40
10.95
1.00
Kennedy et al. 2000
6.90
13.20
20.30
16.75
2.43
Rudolph et al. 2007
7.70
13.90
-
13.90
1.81
Rudolph et al. 2007
8.20
12.50
-
12.50
1.52
Rudolph et al. 2007
8.00
15.00
-
15.00
1.88
Rudolph et al. 2007
8.10
14.90
-
14.90
1.84
Rudolph et al. 2007
6.60
15.20
-
15.20
2.30
Rudolph et al. 2007
8.50
12.90
-
12.90
1.52
Rudolph et al. 2007
7.20
12.30
-
12.30
1.71
Rudolph et al. 2007
7.30
16.70
-
16.70
2.29
Rudolph et al. 2007
7.60
13.10
-
13.10
1.72
Rudolph et al. 2007
8.70
15.60
-
15.60
1.79
Rudolph et al. 2007
8.20
13.90
-
13.90
1.70
Rudolph et al. 2007
7.90
15.10
-
15.10
1.91
Rudolph et al. 2007
7.60
12.30
-
12.30
1.62
Rudolph et al. 2007
11.80
16.10
-
16.10
1.36
Rudolph et al. 2007
40.40
86.30
-
86.30
2.14
Rudolph et al. 2007
46.10
121.00
-
121.00
2.62
Rudolph et al. 2007
50.40
140.00
-
140.00
2.78
Rudolph et al. 2007
34.70
51.00
-
51.00
1.47
Rudolph et al. 2007
39.00
65.30
-
65.30
1.67
Rudolph et al. 2007
35.40
46.80
-
46.80
1.32
Rudolph et al. 2007
11.30
16.90
-
16.90
1.50
Rudolph et al. 2007
13.40
20.60
-
20.60
1.54
150

0
Median ratio:
1.81
100 
/
R2:
0.82
r
^egg-ovary

F:
308.3
50 
y*
df:
P:
67
<0.001
50
100
150
B-28

-------
Dolly Varden (Salvelinus malma)
Study

Cmuscle
c c
^ess v o\ ;tr\
r
v ess o\ ;tr\
Ratio
Golder 2009

73.00
92.30
92.30
1.26
Golder 2009

45.90
40.70
40.70
0.89
Golder 2009

107.00
107.00
107.00
1.00
Golder 2009

97.20
102.00
102.00
1.05
Golder 2009

114.00
124.00
124.00
1.09
Golder 2009

115.00
185.00
185.00
1.61
Golder 2009

79.60
112.00
112.00
1.41
Golder 2009

9.90
7.00
7.00
0.71
Golder 2009

3.40
12.10
12.10
3.56
Golder 2009

5.30
9.60
9.60
1.81
Golder 2009

2.80
5.40
5.40
1.93
Golder 2009

4.90
10.50
10.50
2.14
Golder 2009

6.60
11.00
11.00
1.67
Golder 2009

55.70
65.80
65.80
1.18
Golder 2009

58.30
51.90
51.90
0.89
Golder 2009

39.50
60.50
60.50
1.53
Golder 2009

50.50
56.60
56.60
1.12
200 -|
O




150 


Median ratio:
1.26

100 -
^egg-ovary
50 -
n .
So
y

R2:
F:
df:
0.90
140.3
15

/

P:
<0. 001

0	50	100	150	200
f
^muscle
B-29

-------
Flannelmouth sucker (Catostomus latipinnis)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
3.6

4.0
4.0
1.11
Osmundson et al. 2007
3.8
-
4.1
4.1
1.08
Osmundson et al. 2007
4.1
-
5.9
5.9
1.44
Osmundson et al. 2007
4.6
-
4.3
4.3
0.93
Osmundson et al. 2007
5.2
-
5.7
5.7
1.10
Osmundson et al. 2007
6.2
-
6.2
6.2
1.00
Osmundson et al. 2007
7.3
-
6.2
6.2
0.85
egg-ovary
muscle
Median ratio:	1.08
R2:	0.58
F:	6.92
df:	5
P:	0.036
Green sunfish (Lepomis cyanellus)
Study
r
^muscle
r
egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
28.1

27.4
27.4
0.98
Osmundson et al. 2007
12.9
-
10.2
10.2
0.79
Osmundson et al. 2007
21.9
-
21.8
21.8
1.00
Osmundson et al. 2007
5
-
7
7
1.40
Osmundson et al. 2007
6.1
-
8.9
8.9
1.46
Osmundson et al. 2007
5.2
-
6.4
6.4
1.23
Osmundson et al. 2007
5.1
-
6.4
6.4
1.25
Osmundson et al. 2007
15.7
-
18.1
18.1
1.15
Osmundson et al. 2007
10.1
-
12.3
12.3
1.22
Osmundson et al. 2007
11.5
-
13.8
13.8
1.20
Osmundson et al. 2007
10.5
-
15.2
15.2
1.45
Osmundson et al. 2007
7.2
-
10.8
10.8
1.50
Osmundson et al. 2007
9.3
-
11.7
11.7
1.26
Osmundson et al. 2007
7.7
-
12.6
12.6
1.64
Osmundson et al. 2007
6
-
10
10
1.67
Osmundson et al. 2007
12
-
13.9
13.9
1.16
Osmundson et al. 2007
12.1
-
15.2
15.2
1.26
Osmundson et al. 2007
12.5
-
14.7
14.7
1.18
Osmundson et al. 2007
7.5
-
8.8
8.8
1.17
Osmundson et al. 2007
11.3
-
12.9
12.9
1.14
B-30

-------
Green sunfish (Lepomis cyanellus)
Study
-^muscle
Ratio
Osmundson et al. 2007
13.6
-
13.1
13.1
0.96
Osmundson et al. 2007
13.2
-
11.5
11.5
0.87
Osmundson et al. 2007
12.4
-
13.2
13.2
1.06
Osmundson et al. 2007
12.5
-
11.6
11.6
0.93
Osmundson et al. 2007
8.6
-
7.5
7.5
0.87
Osmundson et al. 2007
5.3
-
8.1
8.1
1.53
Osmundson et al. 2007
11.9
-
13.2
13.2
1.11
Osmundson et al. 2007
13.6
-
14
14
1.03
Osmundson et al. 2007
3.8
-
5.2
5.2
1.37
Osmundson et al. 2007
4.2
-
5.8
5.8
1.38
Osmundson et al. 2007
4.1
-
4.1
4.1
1.00
Osmundson et al. 2007
4.2
-
4.9
4.9
1.17
Osmundson et al. 2007
5.7
-
9.5
9.5
1.67
Osmundson et al. 2007
4.4
-
4.8
4.8
1.09
Osmundson et al. 2007
3.5
-
5.6
5.6
1.60
Osmundson et al. 2007
5.5
-
10.1
10.1
1.84
Osmundson et al. 2007
5
-
7.5
7.5
1.50
Osmundson et al. 2007
4.3
-
5.9
5.9
1.37
30 -I



>	
Median ratio: 1.21
20 





c
egg-ovary

R2: 0.89
10 

F: 281.4


df: 36
0 

P: <0.001
0 10 20 30


c
^muscle

B-31

-------
Largemouth bass (Micropterus salmoicles)
Study
r
^muscle
r
ess
c
*^ovarv
r
^egg-ovarv
Ratio
Carolina Power & Light 1997
8.48

14.79
14.79
1.74
Carolina Power & Light 1997
8.48
-
14.79
14.79
1.74
Carolina Power & Light 1997
7.29
-
8.35
8.35
1.15
Carolina Power & Light 1997
15
-
19
19
1.27
Carolina Power & Light 1997
15
-
15
15
1.00
Carolina Power & Light 1997
12
-
14
14
1.17
Carolina Power & Light 1997
10
-
18
18
1.80
Carolina Power & Light 1997
18
-
15
15
0.83
Carolina Power & Light 1997
18
-
15
15
0.83
Carolina Power & Light 1997
11
-
12
12
1.09
Carolina Power & Light 1997
11
-
9.4
9.4
0.85
Carolina Power & Light 1997
13
-
10
10
0.77
Carolina Power & Light 1997
11
-
11
11
1.00
20
Median ratio: 1.09
R
F
df
P
0.14
1.74
11
0.22
Not used because P>0.05
5	10	15
^muscle
20
B-32

-------
Mountain whitefish (Prosopium williamsoni)
Study
r
^muscle
r
^egg
r
*^ovarv
r
^egg-ovarv
Ratio
Golder 2005
3.60

26.90

26.90
7.47
Golder 2005
3.70
-
25.80

25.80
6.97
Golder 2005
3.10
-
20.00

20.00
6.45
Golder 2005
4.20
-
19.30

19.30
4.60
Golder 2005
3.90
-
19.20

19.20
4.92
Golder 2005
3.50
-
23.20

23.20
6.63
Golder 2005
5.20
-
38.00

38.00
7.31
Golder 2005
5.00
-
41.00

41.00
8.20
Golder 2005
5.20
-
32.00

32.00
6.15
Golder 2005
7.60
-
34.00

34.00
4.47
Golder 2005
7.20
-
32.00

32.00
4.44
Golder 2005
5.50
-
40.00

40.00
7.27
Golder 2005
7.80
-
39.70

39.70
5.09
Golder 2005
3.70
-
20.30

20.30
5.49
Golder 2005
4.70
-
22.40

22.40
4.77
Golder 2005
4.40
-
28.90

28.90
6.57
Golder 2005
5.70
-
30.10

30.10
5.28
Golder 2005
4.00
-
31.50

31.50
7.88
Golder 2005
10.00
-
35.20

35.20
3.52
Golder 2005
4.90
-
26.70

26.70
5.45
Golder 2005
7.60
-
26.80

26.80
3.53
Golder 2005
6.10
-
29.70

29.70
4.87
Golder 2005
6.80
-
41.10

41.10
6.04
Golder 2005
5.00
-
29.00

29.00
5.80
Golder 2005
6.60
-
34.50

34.50
5.23
Golder 2005
5.00
-
36.30

36.30
7.26
Golder 2005
5 -i
4.80
-
28.90

28.90
6.02
s-
Median
ratio:
5.80


p 25 
W
VP o

R2:
0.33


^ egg-ova iy


F:
12.4





df:
25


n


P:
<0.001

0	25	50
r
^muscle
B-33

-------
Northern pike (Esox lucius)
Study
Cmuscle
r
v^egg
c c
^ovarv ^ess-ovarv
Ratio
Muscatello et al. 2006
0.90
3.50
3.50
3.89
Muscatello et al. 2006
1.90
2.70
2.70
1.42
Muscatello et al. 2006
2.60
3.40
3.40
1.31
Muscatello et al. 2006
1.30
3.70
3.70
2.85
Muscatello et al. 2006
1.00
2.70
2.70
2.70
Muscatello et al. 2006
17.00
43.20
43.20
2.54
Muscatello et al. 2006
16.50
24.50
24.50
1.48
Muscatello et al. 2006
16.50
26.10
26.10
1.58
Muscatello et al. 2006
2.00
3.40
3.40
1.70
Muscatello et al. 2006
2.00
4.10
4.10
2.05
Muscatello et al. 2006
1.30
4.10
4.10
3.15
Muscatello et al. 2006
2.50
4.10
4.10
1.64
Muscatello et al. 2006
1.30
3.40
3.40
2.62
Muscatello et al. 2006
47.80
48.20
48.20
1.01
60
30 -
30
p
^muscle
60
Median ratio: 1 .<
R2:
0.83
F:
58.9
df:
12
P:
<0.001
Rainbow trout (Oncorhynchus mykiss)
Study
C nmscle
r
v^egg
c c
^ovarv ^ess-ovarv
Ratio
Casey and Siwik 2000
4.10
11.60
11.60
2.83
Casey and Siwik 2000
3.80
10.10
10.10
2.66
Casey and Siwik 2000
2.60
0.10
0.10
0.04
Casey and Siwik 2000
3.30
4.90
4.90
1.48
Casey and Siwik 2000
2.30
3.60
3.60
1.57
Casey and Siwik 2000
2.80
5.30
5.30
1.89
Casey and Siwik 2000
2.30
3.70
3.70
1.61
Casey and Siwik 2000
2.80
6.40
6.40
2.29
Casey and Siwik 2000
3.00
5.20
5.20
1.73
Casey and Siwik 2000
4.90
6.80
6.80
1.39
Casey and Siwik 2000
1.50
3.60
3.60
2.40
Casey and Siwik 2000
2.60
6.90
6.90
2.65
B-34

-------
Rainbow trout (Oncorhynchus mykiss)
Study
C nmscle
r
v^egg
c c
^ovarv ^ess-ovarv
Ratio
Casey and Siwik 2000
4.60
6.90
6.90
1.50
Casey and Siwik 2000
4.60
6.40
6.40
1.39
Casey and Siwik 2000
3.60
5.50
5.50
1.53
Casey and Siwik 2000
2.40
10.50
10.50
4.38
Casey and Siwik 2000
3.70
7.60
7.60
2.05
Casey and Siwik 2000
2.70
4.10
4.10
1.52
Casey and Siwik 2000
0.70
1.10
1.10
1.57
Casey and Siwik 2000
0.60
0.90
0.90
1.50
Casey and Siwik 2000
0.60
1.30
1.30
2.17
Casey and Siwik 2000
28.60
56.30
56.30
1.97
Casey and Siwik 2000
30.90
56.00
56.00
1.81
Casey and Siwik 2000
32.40
71.50
71.50
2.21
Casey and Siwik 2000
28.00
61.30
61.30
2.19
Casey and Siwik 2000
31.70
54.50
54.50
1.72
Casey and Siwik 2000
29.50
56.80
56.80
1.93
Casey and Siwik 2000
30.10
57.90
57.90
1.92
Casey and Siwik 2000
29.90
64.70
64.70
2.16
Casey and Siwik 2000
32.80
46.60
46.60
1.42
Casey and Siwik 2000
31.40
56.50
56.50
1.80
Casey and Siwik 2000
32.00
67.50
67.50
2.11
Casey and Siwik 2000
35.70
59.40
59.40
1.66
Casey and Siwik 2000
24.60
48.70
48.70
1.98
Casey and Siwik 2000
30.30
69.10
69.10
2.28
Casey and Siwik 2000
25.70
43.50
43.50
1.69
Casey and Siwik 2000
35.00
58.10
58.10
1.66
Casey and Siwik 2000
33.80
59.20
59.20
1.75
Casey and Siwik 2000
28.70
55.00
55.00
1.92
Casey and Siwik 2000
25.80
49.00
49.00
1.90
Holm et al. 2005
1.70
1.00
1.00
0.59
Holm et al. 2005
1.60
3.50
3.50
2.19
Holm et al. 2005
1.30
4.60
4.60
3.54
Holm et al. 2005
4.00
12.80
12.80
3.20
Holm et al. 2005
4.30
17.10
17.10
3.98
Holm et al. 2005
8.50
17.50
17.50
2.06
Holm et al. 2005
7.40
29.70
29.70
4.01
B-35

-------
Rainbow trout (Oncorhynchus mykiss)
Ratio
muscle
ovary
egg-ovary
Median ratio
egg-ovary
muscle
Razorback sucker (Xyrauchen texanus)
Study
Cmuscle
r
v^egg
C C
^ovarv ^egg-ovarv
Ratio
Hamilton et al. 2001a
5
7.5
7.5
1.50
Hamilton et al. 2001a
4
6.1
6.1
1.53
Hamilton et al. 2001a
4.2
6.6
6.6
1.57
Hamilton et al. 2001a
4.4
6.2
6.2
1.41
Hamilton et al. 2001a
4.5
5.8
5.8
1.29
Hamilton et al. 2001a
4.3
6.8
6.8
1.58
Hamilton et al. 2001a
11.1
35.5
35.5
3.20
Hamilton et al. 2001a
12.2
43.4
43.4
3.56
Hamilton et al. 2001a
10.4
54.5
54.5
5.24
Hamilton et al. 2001a
11.3
28.2
28.2
2.50
Hamilton et al. 2001a
10.4
38
38
3.65
Hamilton et al. 2001a
17.3
41.3
41.3
2.39
Hamilton et al. 2001a
13
37.2
37.2
2.86
Hamilton et al. 2001a
16.7
40.9
40.9
2.45
Hamilton et al. 2001a
14.6
35.3
35.3
2.42
Hamilton et al. 2001a
12.1
34.3
34.3
2.83
Hamilton et al. 2001b
4.7
5
5
1.06
Hamilton et al. 2001b
5.3
6.2
6.2
1.17
Hamilton et al. 2001b
3.6
5.9
5.9
1.64
Hamilton et al. 2001b
5.3
6.5
6.5
1.23
Hamilton et al. 2001b
4.1
6.35
6.35
1.55
Hamilton et al. 2001b
4.9
6.1
6.1
1.24
Hamilton et al. 2001b
16
40.1
40.1
2.51
Hamilton et al. 2001b
18
38.4
38.4
2.13
Hamilton et al. 2001b
16
40.2
40.2
2.51
Hamilton et al. 2001b
19
43.1
43.1
2.27
B-36

-------
Razorback sucker (Xyrauchen texanus)
Study
Cmuscle
r
v^egg
c c
^ovarv ^ess-ovarv
Ratio
Hamilton et al. 2001b
14
41.9
41.9
2.99
Hamilton et al. 2001b
14
36.2
36.2
2.59
Hamilton et al. 2001b
24
56.5
56.5
2.35
Hamilton et al. 2001b
27
51.8
51.8
1.92
Hamilton et al. 2001b
24
52.6
52.6
2.19
Hamilton et al. 2001b
27
55.1
55.1
2.04
Hamilton et al. 2001b
19
53
53
2.79
Hamilton et al. 2001b
16
58.5
58.5
3.66
Waddell and May 1995 a
4.40
3.70
3.70
X
Waddell and May 1995 a
7.10
4.70
4.70
X
Waddell and May 1995 a
32.00
10.60
10.60
X
70
60
50
40 -
30 
20 
10 
Q/lg
/
L

0
10
20
30 40
c
^DIQSCie
50
60
70
Median ratio: 2.31
R2:
0.80
F:
125.6
df:
32
P:
<0.001
Data from this study were excluded because results were atypical.
Roundtail chub (Gila robusta)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
4.3

7.9
7.9
1.84
Osmundson et al. 2007
5
-
10.8
10.8
2.16
Osmundson et al. 2007
6.2
-
15.2
15.2
2.45
Osmundson et al. 2007
6.9
-

14.1
2.04
Osmundson et al. 2007
7
-
10.6
10.6
1.51
Osmundson et al. 2007
7.3
-
18
18
2.47
Osmundson et al. 2007
9.8
-
17.8
17.8
1.82
B-37

-------
Roundtail chub (Gila robusta)
ovary
muscle
Median ratio: 2.04
R2:
F:
df:
P:
0.62
8.27
5
0.026
Smallmouth bass (Micropterus dolomieu)
Study
r
^muscle
r
ess
c
*^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
3.7

6.0
6.0
1.62
Osmundson et al. 2007
6.5
-
8.0
8.0
1.23
Osmundson et al. 2007
6.9
-
6.5
6.5
0.94
Osmundson et al. 2007
11


11
1.00
Osmundson et al. 2007
5.5
-
7.1
7.1
1.29
Osmundson et al. 2007
7.7
-
8.8
8.8
1.14
15
10
'egg-ovary
5	10
f
^muscle
15
Median ratio:
1.19
R2:
0.85
F:
23.5
df:
4
P:
0.006
B-38

-------
White Sturgeon (Acipenser transmontanus)
Study
Cmuscle
r
v^egg
C C
^ovarv ^ess-ovarv
Ratio
Linville 2006
1.28
2.46
2.46
2.46
Linville 2006
1.22
1.61
1.61
1.61
Linville 2006
1.48
2.68
2.68
2.68
Linville 2006
9.93
11
11
11
Linville 2006
15.3
20.5
20.5
20.5
Linville 2006
11.1
7.61
7.61
7.61
10
c
v muscle
Median ratio: 1.33
R2:
0.86
F:
24.96
df:
4
P:
0.006
White Sucker (Catostomus commersonii)
Study
r
^muscle
r
egg
c
*^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
2.9

6.2
6.2
2.14
Osmundson et al. 2007
4.8
-
6.2
6.2
1.29
Osmundson et al. 2007
3.7
-
5.2
5.2
1.41
Osmundson et al. 2007
3.7
-
6.5
6.5
1.76
Osmundson et al. 2007
8.4
-
7.7
7.7
0.92
Osmundson et al. 2007
9.4
-
5.8
5.8
0.62
Osmundson et al. 2007
15.5
-
10.9
10.9
0.70
Osmundson et al. 2007
23.6
-
11.2
11.2
0.47
Osmundson et al. 2007
9.4
-
9.4
9.4
1.00
Osmundson et al. 2007
6.1
-
5.4
5.4
0.89
Osmundson et al. 2007
4.6
-
5.1
5.1
1.11
Osmundson et al. 2007
12.3
-
10.4
10.4
0.85
Osmundson et al. 2007
9.2
-
10.4
10.4
1.13
Osmundson et al. 2007
9.4
-
11
11
1.17
Osmundson et al. 2007
9.4
-
11.7
11.7
1.24
Osmundson et al. 2007
10.5
-
11.6
11.6
1.10
Osmundson et al. 2007
11.4
-
9.4
9.4
0.82
Osmundson et al. 2007
9.6
-
10.2
10.2
1.06
Osmundson et al. 2007
9.3
-
7.3
7.3
0.78
Osmundson et al. 2007
9.8
-
8.9
8.9
0.91
Osmundson et al. 2007
10.5
-
10.5
10.5
1.00
B-39

-------
White Sucker (Catostomus commersonii)
Study
Cmuscle
r
v^egg
r
^ovarv
r
^ess-ovarv
Ratio
Osmundson et al. 2007
11.1

10.2
10.2
0.92
Osmundson et al. 2007
12.1
-
8.1
8.1
0.67
Osmundson et al. 2007
12.8
-
9.5
9.5
0.74
Osmundson et al. 2007
16.0
-
10.7
10.7
0.67
Osmundson et al. 2007
12.1
-
8.3
8.3
0.69
Osmundson et al. 2007
9.0
-
12
12
1.33
Osmundson et al. 2007
10.6
-
6.1
6.1
0.58
Osmundson et al. 2007
12.6
-
6.1
6.1
0.48
Osmundson et al. 2007
11.6
-
11.3
11.3
0.97
Osmundson et al. 2007
2.8
-
2.6
2.6
0.93
Osmundson et al. 2007
2.5
-
3.6
3.6
1.44
Osmundson et al. 2007
4.3
-
4.4
4.4
1.02
Osmundson et al. 2007
3.5
-
4.4
4.4
1.26
Osmundson et al. 2007
4.3
-
4.8
4.8
1.12
Osmundson et al. 2007
3.1
-
4.3
4.3
1.39
Osmundson et al. 2007
3.6
-
4.1
4.1
1.14
Osmundson et al. 2007
3.0
-
3.8
3.8
1.27
Osmundson et al. 2007
4.1
-
3.6
3.6
0.88
Osmundson et al. 2007
3.6
-
3.8
3.8
1.06
egg-ovary
muscle
Median ratio:	1.00
R2: 0.59
F:	53.92
df:	38
P:	<0.001
B-40

-------
Table B-3. Summary of egg-ovary to muscle conversion factors.
Common name
Scientific name
Median ratio
Bluegill
Lepomis macrochirus
1.38
Bluehead sucker
Catostomus discobolus
1.48
Brook trout
Salvelinus fontinalis
1.09
Common carp
Cyprinus carpio
1.14
Cutthroat trout
Oncorhynchus clarkii
1.81
Dolly Varden
Salvelinus malma
1.26
Flannelmouth sucker
Catostomus latipinnis
1.08
Green sunfish
Lepomis cyanellus
1.21
Mountain whitefish
Prosopium williamsoni
5.80
Northern pike
Esox lucius
1.88
Rainbow trout
Oncorhynchus mykiss
1.92
Razorback sucker
Xyrauchen texanus
2.31
Roundtail chub
Gila robusta
2.04
Smallmouth bass
Micropterus dolomieu
1.19
White sturgeon
Acipenser transmontanus
1.33
White sucker
Catostomus commersonii
1.00
B-41

-------
2.4 Muscle to whole-body conversion factors
n
^ whole-body
^muscle
Ratio
Selenium concentration in all tissues ((.ig/g dw)
Selenium concentration in muscle tissue only (fj.g/g dw)
c.
x muscle
C
whole-body
Black bullhead (Ameiurus melas)
Study


Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007

5.30
3.40
0.64
Osmundson et al. 2007

4.80
3.90
0.81
Osmundson et al. 2007

5.50
4.30
0.78
Osmundson et al. 2007

4.90
4.70
0.96
Osmundson et al. 2007

9.60
5.70
0.59
Osmundson et al. 2007

7.60
7.40
0.97
Osmundson et al. 2007

7.30
7.50
1.03
Osmundson et al. 2007

6.60
7.80
1.18
Osmundson et al. 2007

8.60
7.80
0.91
Osmundson et a
15 1
[. 2007

2.00
9.20
4.60
10 


Median ratio:
0.93
f
^muscle
5 
o
OQ O

R2:
0.00

o

F:
0.03


o

df:
8
n



P:
0.973
u

5 10
15 Not used because P > 0.05.
-'whole-body
Bluegill (Lepomis macrochirus)
Study
Cwhole-bodv
Cmuscle
Ratio
Doroshov et al. 1992
1.60
1.50
0.94
Doroshov et al. 1992
5.50
5.80
1.05
Doroshov et al. 1992
9.30
10.40
1.12
Doroshov et al. 1992
19.30
23.60
1.22
Hermanutz et al. 1996
1.50
2.10
1.40
Hermanutz et al. 1996
18.10
20.60
1.14
Hermanutz et al. 1996
1.90
1.90
1.00
Hermanutz et al. 1996
2.80
3.50
1.25
B-42

-------
Bluegill (Lepomis macrochirus)
Study

Cwhole-bodv
Cmuscle
Ratio
Hermanutz et al.
996
12.30
17.60
1.43
Hermanutz et al.
996
9.40
12.50
1.33
Hermanutz et al.
996
1.50
2.30
1.53
Hermanutz et al.
996
4.90
6.90
1.41
Hermanutz et al.
996
21.00
44.90
2.14
Hermanutz et al.
996
24.30
39.80
1.64
Hermanutz et al.
996
2.70
3.40
1.26
Hermanutz et al.
996
5.00
5.30
1.06
Hermanutz et al.
996
9.50
12.50
1.32
Hermanutz et al.
996
6.60
7.80
1.18
Hermanutz et al.
996
1.80
3.20
1.78
Hermanutz et al.
996
4.20
6.10
1.45
Hermanutz et al.
996
10.30
18.70
1.82
Hermanutz et al.
996
13.80
15.10
1.09
Osmundson et al.
2007
8.80
12.90
1.47
-muscle
50
25
Median ratio: 1.32
R
F
df
P
0.89
172.2
21
<0.001
25
50
-whole-body
B-43

-------
Bluehead sucker (Catostomus discobolus)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
1.30
1.50
1.15
Osmundson et al. 2007
2.00
2.30
1.15
Osmundson et al. 2007
2.10
2.50
1.19
Osmundson et al. 2007
2.20
2.70
1.23
Osmundson et al. 2007
2.40
3.10
1.29
Osmundson et al. 2007
3.90
5.20
1.33
Osmundson et al. 2007
5.60
8.60
1.54
10 n
-muscle
Median ratio: 1.23
R
F
df
P
10
"whole-body
0.99
682.9
5
<0.001
Brown trout (Salmo trutta)
Study
whole-bodv
muscle
Ratio
Osmundson et al. 2007
4.60
3.20
0.70
Osmundson et al. 2007
4.30
3.60
0.84
Osmundson et al. 2007
5.00
4.00
0.80
Osmundson et al. 2007
5.50
6.30
1.15
4 -
Median ratio: 0.82
R2:
0.78
F:
7.2
df:
2
P:
0.122
Not used because P > 0.05.
"whole-body
B-44

-------
Channel catfish (Ictalurus punctatus)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
3.40
3.40
1.00
Osmundson et al. 2007
3.30
3.60
1.09
Osmundson et al. 2007
2.60
3.70
1.42
Osmundson et al. 2007
4.00
5.30
1.33
"muscle
%
Median ratio: 1.21
R2:
0.49
F:
2.0
df:
2
P:
0.338
Not used because P > 0.05.
"whole-body
Common carp (Cyprinus carpio)
Study
r1
whole-bodv
Cmuscle Rati O

Osmundson et al. 2007
6.30
7.80
1.24
Osmundson et al. 2007
4.80
8.20
1.71
Osmundson et al. 2007
11.70
20.00
1.71
Osmundson et al. 2007
23.10
24.20
1.05
Osmundson et al. 2007
4.10
6.60
1.61
30
20
10 
O o
10
20
30
Median ratio: 1.61
R2:
0.85
F:
17.6
df:
3
P:
0.017
-whole-body
B-45

-------
Flannelmouth sucker (Catostomus latipinnis)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
3.0
3.6
1.20
Osmundson et al. 2007
2.6
3.8
1.46
Osmundson et al. 2007
3.1
4.1
1.32
Osmundson et al. 2007
3.1
4.6
1.48
Osmundson et al. 2007
3.5
5.2
1.49
Osmundson et al. 2007
4.4
6.2
1.41
Osmundson et al. 2007
4.5
7.3
1.62
J
C	5
^muscle
whole body
Median ratio: 1.46
R2:
0.91
F:
50.1
df:
5
P:
<0.001
Green sunfish (Lepomis cyanellus)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
22.80
28.10
1.23
Osmundson et al. 2007
8.80
12.90
1.47
Osmundson et al. 2007
15.40
21.90
1.42
Osmundson et al. 2007
4.80
5.00
1.04
Osmundson et al. 2007
5.70
6.10
1.07
Osmundson et al. 2007
4.40
5.20
1.18
Osmundson et al. 2007
3.80
5.10
1.34
Osmundson et al. 2007
11.90
15.70
1.32
Osmundson et al. 2007
6.40
10.10
1.58
Osmundson et al. 2007
9.50
11.50
1.21
Osmundson et al. 2007
9.10
10.50
1.15
Osmundson et al. 2007
6.20
7.20
1.16
Osmundson et al. 2007
7.00
9.30
1.33
Osmundson et al. 2007
7.70
7.70
1.00
Osmundson et al. 2007
6.20
6.00
0.97
Osmundson et al. 2007
10.20
12.00
1.18
Osmundson et al. 2007
9.70
12.10
1.25
Osmundson et al. 2007
9.90
12.50
1.26
Osmundson et al. 2007
7.20
7.50
1.04
Osmundson et al. 2007
9.00
11.30
1.26
B-46

-------
Green sunfish (Lepomis cyanellus)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
9.70
13.60
1.40
Osmundson et al. 2007
8.90
13.20
1.48
Osmundson et al. 2007
9.80
12.40
1.27
Osmundson et al. 2007
9.90
12.50
1.26
Osmundson et al. 2007
10.30
8.60
0.83
Osmundson et al. 2007
5.30
5.30
1.00
Osmundson et al. 2007
10.10
11.90
1.18
Osmundson et al. 2007
11.80
13.60
1.15
Osmundson et al. 2007
3.30
3.80
1.15
Osmundson et al. 2007
4.00
4.20
1.05
Osmundson et al. 2007
4.30
4.10
0.95
Osmundson et al. 2007
3.70
4.20
1.14
Osmundson et al. 2007
6.20
5.70
0.92
Osmundson et al. 2007
3.50
4.40
1.26
Osmundson et al. 2007
4.40
3.50
0.80
Osmundson et al. 2007
5.60
5.50
0.98
Osmundson et al. 2007
4.90
5.00
1.02
Osmundson et al. 2007
4.40
4.30
0.98
Osmundson et al. 2007
8.00
10.10
1.26
Osmundson et al. 2007
7.90
11.90
1.51
Osmundson et al. 2007
6.40
11.10
1.73
Osmundson et al. 2007
8.70
11.80
1.36
Osmundson et al. 2007
8.30
11.00
1.33
Osmundson et al. 2007
6.10
7.10
1.16
Osmundson et al. 2007
5.60
6.70
1.20
Osmundson et al. 2007
18.10
26.40
1.46
Osmundson et al. 2007
9.40
9.60
1.02
Osmundson et al. 2007
12.20
16.70
1.37
Osmundson et al. 2007
5.30
8.10
1.53
Osmundson et al. 2007
7.30
10.60
1.45
Osmundson et al. 2007
9.30
14.20
1.53
Osmundson et al. 2007
6.80
11.30
1.66
Osmundson et al. 2007
7.50
12.80
1.71
B-47

-------
Green sunfish (Lepomis cyanellus)
Ratio
whole-body
muscle
Median ratio
muscle
whole-body
Roundtail chub (Gila robusta)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
4.10
4.30
1.05
Osmundson et al. 2007
5.30
5.00
0.94
Osmundson et al. 2007
6.40
6.20
0.97
Osmundson et al. 2007
6.80
6.90
1.01
Osmundson et al. 2007
5.50
7.00
1.27
Osmundson et al. 2007
6.60
7.30
1.11
Osmundson et al. 2007
8.40
9.80
1.17
Median ratio
muscle
whole-body
B-48

-------
Smallmouth bass (Micropterus dolomieu)
Study	
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
Osmundson et al. 2007
12 -i
o
0 -I	1	
0	6
f
^whole-body
Cwhole-bodv
Cmuscle
Ratio
4.20
3.70
0.88
5.50
6.50
1.18
5.40
6.90
1.28
7.80
11.0
1.41
5.10
7.10
1.08
4.90
8.80
1.57
Median ratio: 1.23
R2:
0.83
F:
20.2
df:
4
P:
0.008
White sucker (Catostomus commersonii)
Study
^whole-bodv
^ muscle
Ratio
Osmundson et al. 2007
3.80
2.90
0.76
Osmundson et al. 2007
4.20
4.80
1.14
Osmundson et al. 2007
3.30
3.70
1.12
Osmundson et al. 2007
4.50
3.70
0.82
Osmundson et al. 2007
6.30
8.40
1.33
Osmundson et al. 2007
6.80
9.40
1.38
Osmundson et al. 2007
11.00
15.50
1.41
Osmundson et al. 2007
12.70
23.60
1.86
Osmundson et al. 2007
5.70
9.40
1.65
Osmundson et al. 2007
3.90
6.10
1.56
Osmundson et al. 2007
3.80
4.60
1.21
Osmundson et al. 2007
9.90
12.30
1.24
Osmundson et al. 2007
5.30
9.20
1.74
Osmundson et al. 2007
10.70
9.40
0.88
Osmundson et al. 2007
5.90
9.40
1.59
Osmundson et al. 2007
7.00
10.50
1.50
Osmundson et al. 2007
6.40
11.40
1.78
Osmundson et al. 2007
6.30
9.60
1.52
Osmundson et al. 2007
5.30
9.30
1.75
Osmundson et al. 2007
6.20
9.80
1.58
B-49

-------
White sucker (Catostomus commersonii)
Study
Cwhole-bodv
Cmuscle
Ratio
Osmundson et al. 2007
5.60
10.50
1.88
Osmundson et al. 2007
8.80
11.10
1.26
Osmundson et al. 2007
8.70
12.10
1.39
Osmundson et al. 2007
11.40
12.80
1.12
Osmundson et al. 2007
10.70
16.00
1.50
Osmundson et al. 2007
8.40
12.10
1.44
Osmundson et al. 2007
7.00
9.00
1.29
Osmundson et al. 2007
7.50
10.60
1.41
Osmundson et al. 2007
10.30
12.60
1.22
Osmundson et al. 2007
6.70
11.60
1.73
Osmundson et al. 2007
2.10
2.80
1.33
Osmundson et al. 2007
1.80
2.50
1.39
Osmundson et al. 2007
3.20
4.30
1.34
Osmundson et al. 2007
2.30
3.50
1.52
Osmundson et al. 2007
3.10
4.30
1.39
Osmundson et al. 2007
3.00
3.10
1.03
Osmundson et al. 2007
2.80
3.60
1.29
Osmundson et al. 2007
2.50
3.00
1.20
Osmundson et al. 2007
3.40
4.10
1.21
Osmundson et al. 2007
2.80
3.60
1.29
Osmundson et al. 2007
3.10
5.60
1.81
Osmundson et al. 2007
5.50
6.30
1.15
Osmundson et al. 2007
7.00
9.10
1.30
Osmundson et al. 2007
7.30
8.50
1.16
Osmundson et al. 2007
2.40
3.00
1.25
Osmundson et al. 2007
2.70
4.40
1.63
Osmundson et al. 2007
2.70
3.20
1.19
Osmundson et al. 2007
2.60
1.60
0.62
Osmundson et al. 2007
19.60
28.10
1.43
Osmundson et al. 2007
9.80
12.10
1.23
Osmundson et al. 2007
8.70
11.80
1.36
Osmundson et al. 2007
8.70
12.60
1.45
Osmundson et al. 2007
9.10
12.30
1.35
Osmundson et al. 2007
13.40
18.00
1.34
Osmundson et al. 2007
3.10
2.80
0.90
Osmundson et al. 2007
2.40
3.20
1.33
Osmundson et al. 2007
2.10
3.10
1.48
Osmundson et al. 2007
3.20
4.30
1.34
Osmundson et al. 2007
2.80
3.40
1.21
B-50

-------
White sucker (Catostomus commersonii)
Ratio
whole-body
muscle
Median ratio
muscle
whole-body
Table B-4. Muscle to whole-body correction factor.
(0111111011 name
Scientific name
Median ratio
Bluegill
Lepomis macrochirus
1.32
Bluehead sucker
Catostomus discobolus
1.23
Common carp
Cyprinus carpio
1.61
Flannelmouth sucker
Catostomus latipinnis
1.46
Green sunfish
Lepomis cyanellus
1.23
Roundtail chub
Gila robusta
1.05
Smallmouth bass
Micropterus dolomieu
1.23
White sucker
Catostomus commersonii
1.34
Table B-5. Directly calculated final egg-ovary to whole body conversion factors (CF).
Common name
Median ratio
( eax-oYtm^ ^ whote-hoch")
Median ratio
( ^ eiy>-ovar\^ ^ muscle)
Muscle to
whole-body
correction
factor
l-'inal (7
\ allies
Species
Bluegill
2.13


2.13
Bluehead sucker
1.82


1.82
Brook trout

1.09
1.27
1.38
Brown trout
1.45


1.45
Common carp
1.92


1.92
Creek chub
1.99


1.99
Cutthroat trout
1.96


1.96
B-51

-------
Common 11 :imo
Modi;tii ratio
(Ceee_
Median ratio
( ^ exx-ovary muscle)
Muscle to
whole-body
correction
factor
1 inal(/
\ allies
Desert pupfish
1.20


1.20
Dolly Varden

1.26
1.27
1.61
Fathead minnow
1.40


1.40
Flannelmouth sucker
1.41


1.41
Green sunfish
1.45


1.45
Mountain whitefish

5.80
1.27
7.39
Northern pike

1.88
1.27
2.39
Rainbow trout

1.92
1.27
2.44
Razorback sucker

2.31
1.34
3.11
Roundtail chub
2.07


2.07
Smallmouth bass
1.42


1.42
White sturgeon

1.33
1.27
1.69
White sucker
1.38


1.38

Genus
Catostomus



1.41
Gila



2.07
Lepomis



1.79
Micropterus



1.42
Oncorhynchus



1.96

Family
Catostomidae



1.41
Centrarchidae



1.45
Cyprinidae



1.95
Salmonidae



1.71

Order
Cyprinodontiformes



1.20
Perciformes



1.45
B-52

-------
Common 11 :imo
Modi;tii ratio
( L'xx-uvurx^ whole-bath")
Median ratio
( ^ L'UX-fH'an ^-muscle)
Muscle to
whole-body
correction
factor
Tinal CI'
\ allies

Class
Actinopterygii



1.45
B-53

-------
Table B-6. All EPA-derived egg-ovary to whole body (CF), egg-ovary to muscle, and muscle to whole body conversion factors directly calculated or estimated using taxonomic classification.
(See main text for explanation of the taxonomic classification approach).


Diivii i;iluil;ili;im-(I on l;i\(iniimii il;issil'ii;ili
-------


Diri-il i;iliul;ili;isi-tl mi l;i\imnmii il;issil'ii;ili
-------
( <11111111111
11.11111-
Si'ii'lllilli' 11 ;i UK-
Diivii i;iluil;ili;isi-il on l;i\(iniimii il;issil'ii;ili
-------


Diivii i;iluil;ili
-------


Diivii i;iluil;ili
-------
3.0 Derivation of Trophic Transfer Factor Values
3.1 Methodology
Taxa specific trophic transfer factors (TTF) to quantify the degree of biomagnification across a
given trophic level were calculated from either physiological parameters measured in laboratory studies
or from field measurements of paired selenium concentrations in consumer species and their food. TTFs
from both approaches were used to calculate translated water concentrations; however, when TTF data of
similar quality are available from both approached, as was the case with bluegill, field-derived TTF data
are used.
Physiological data consisted of assimilation efficiencies (AE), measured as either a percentage or
a proportion, ingestion rates (IR), measured as grams of Se per grams of food consumed per day, and
efflux rate constant (ke), measured as 1/day. All available data were collected for a particular species, and
then the TTF for that species was calculated using the equation:
 AE X IR
TTF =	
ke
Where AE, IR, and Ke were estimated as the median value of all available data for that parameter for that
species.
The majority of TTF were calculated using paired whole-body Se measurements from organisms
collected at the same site in the field. TTFs for trophic level 2 organisms were determined using the
equation:
rT.L2
mm t?T 1/2 	 tissue
111
nTLZ
^food.
Where Cj^d equals the average Se concentration in particulate matter, defined as the average of Caigae,
C detritus? and Csediment- Of the three types of particulate matter potentially assumed by TL2 organisms (e.g.,
the majority of invertebrates), Csediment correlated relatively poorly to Cj^ue, when compared to Caigae and
Cdetritus  In order to minimize potentially erroneous TTF calculations based solely on sediment Se
concentrations, while note completely discounting the importance of organic matter in sediments as a
potential food source, Csediment was included in Cpartlcuiate calculations only when either Caigae or Cdetntus data
were also available.
B-59

-------
TTFs for trophic level 3 organisms were determined using the equation:
rTL3
tt uTL3 	 tissue
1 1 r  vTo
flLi
^food
Where Cj^d equals the average whole-body Se concentration in invertebrates collected at the same site
as their potential predator species. The majority of trophic level 3 organisms are fish species, but
damselflies and dragonflies of the order Odonata are also trophic level 3 organisms, and TTFTL3 values
were calculated for those species as well.
For all field derived data used to determine TTFs, EPA first confirmed a statistical relationship
between whole-body selenium concentrations for each species and its food using OLS linear regression. If
the regression resulted in a statistically significant (P<0.05) positive slope, EPA calculated the TTF as the
median ratio of the paired concentration data.
B-60

-------
3.2 TTF values from physiological coefficients
AE (%) = Assimilation efficiency
IR (g g"1 d"1) = Ingestion rate
ke(d_1)	= Efflux rate constant
jjp	_	AExIR
KR
3.2.1 Invertebrates
Baltic macoma (Macoma balthica)
Physiological Parameters
AE (%) IR(g g1 d ')
ke (d"1)
TTF Study
22.5
91.0
84.0
95.0
78.0	0.03
74.0	0.03
92.3
58.0
85.8
64.9
90.4
Median Values and TTF
84.0	0.27a 0.03
Luomaetal. 1992
Luomaetal. 1992
Luomaetal. 1992
Luomaetal. 1992
Reinfelder et al. 1997
Reinfelder et al. 1997
Schleckat et al. 2002
Schleckat et al. 2002
Schleckat et al. 2002
Schleckat et al. 2002
Schleckat et al. 2002
7.56
' Value taken from Mytilus edulis
Short-necked clam (Ruditapes philippinarum)
Physiological Parameters
AE (%) IR(g g1 d ')
ke (d_1) TTF Study
70.0
52.0
Median Values and TTF
61.0	0.27a
0.013
0.013
0.013 12.67
Zhang et al. 1990
Zhang et al. 1990
' Value taken from Mytilus edulis
B-61

-------
Quahog (Mercenaria mercenaria)	
Physiological Parameters
AE (%) IR(g g1 d ') ke (d1) TTF Study	
Reinfelder and Fisher
100.1	1994
92.0	0.01	Reinfelder et al. 1997
Median Values and TTF
96.1	0.27a	0.01 25.93
a Value taken from Mytilus edulis
Eastern Oyster (Crassostrea virginica)	
Physiological Parameters
AE (%) IR(g g1 d ') ke (d1) TTF Study	
Okazaki and Panietz
0.005	1981
Reinfelder and Fisher
105.4	1994
70.0	0.070	Reinfelder et al. 1997
Median Values and TTF
87.7	0.27a 0.038 6.31
a Value taken from Mytilus edulis
Common mussel (Mytilus edulis)	
Physiological Parameters
AE (%)
IR(g g1 d"1)
ke (d"1)
TTF Study
86.0

0.02
Reinfelder et al. 1997
75.0

0.05
Reinfelder et al. 1997
60.7


Wang and Fisher 1996
48.0


Wang and Fisher 1996
13.7


Wang and Fisher 1996
55.1


Wang and Fisher 1996
55.8


Wang and Fisher 1996
71.9


Wang and Fisher 1996
71.5


Wang and Fisher 1996
27.9


Wang and Fisher 1996
84.4


Wang and Fisher 1996
81.0


Wang and Fisher 1996
79.4


Wang and Fisher 1996
B-62

-------
Physiological Parameters
AE (%) IR(g g1 dx) ke (dx) TTF Study
63.0
0.037
Wang and Fisher 1996
61.5
0.05
Wang and Fisher 1996
69.0
0.027
Wang and Fisher 1996
81.0
0.022
Wang and Fisher 1997
82.0
0.020
Wang and Fisher 1997
72.0
0.018
Wang and Fisher 1997
78.0
0.055
Wang et al. 1995
76.0
0.065
Wang et al. 1995
71.0
0.058
Wang et al. 1995
33.9

Wang et al. 1996
27.5

Wang et al. 1996


Wang et al. 1996


Wang et al. 1996

0.27 0.022
Wang et al. 1996

0.026
Wang et al. 1996

0.019
Wang et al. 1996
Median Values and TTF
71.3	0.27 0.026 7.30
Asian clam (Corbicula fluminea)
Physiological Parameters


AE (%) IR(g g1 d ')
ke (d"1)
TTF Study
55.0 0.05
0.006
Lee et al. 2006
Median Values and TTF


55.0 0.05
0.006
4.58
B-63

-------
Zebra mussel (Dreissena polymorpha)
Physiological Parameters
AE (%)
IR(g g1 d"1)
ke (d"1)
TTF Study
18.0


Roditi and Fisher 1999
24.0


Roditi and Fisher 1999
46.0


Roditi and Fisher 1999
40.0


Roditi and Fisher 1999
41.0


Roditi and Fisher 1999
7.7


Roditi and Fisher 1999
23.0


Roditi and Fisher 1999
28.0
0.40
0.026
Roditi and Fisher 1999
Roditi and Fisher 1999
Roditi and Fisher 1999
Median Values and TTF
26.0	0.40 0.026 4.00
Water flea (Daphnia magna)	
Physiological Parameters
AE (%) IR(g g1 d ') ke (d1) TTF Study

0.08

Goulet et al. 2007

0.34

Goulet et al. 2007
57.9


Yu and Wang 2002b
43.0


Yu and Wang 2002b
39.8


Yu and Wang 2002b
33.0


Yu and Wang 2002b
41.4


Yu and Wang 2002b
41.5


Yu and Wang 2002b
38.0


Yu and Wang 2002b
24.5


Yu and Wang 2002b


0.101
Yu and Wang 2002b


0.12
Yu and Wang 2002b


0.131
Yu and Wang 2002b


0.134
Yu and Wang 2002b


0.108
Yu and Wang 2002b


0.112
Yu and Wang 2002b
Median Values and TTF
40.6	0.21 0.12 0.74
B-64

-------
Copepod (Temora longicornis)
Physiological Parameters
AE (%) IR(g g1 d_1) ke (d') TTF Study
55.0 0.42
0.115

Wang and Fisher 1998
Median Values and TTF



55.0 0.42
0.115
2.01
Copepod (Small, unidentified)
Physiological Parameters
AE (%) IR(g g1 d ')
ke (d"1)
TTF
Study
50.0 0.42
0.155

Schlekat et al. 2004
Median Values and TTF



50.0 0.42
0.155
1.35

Copepod (Large, unidentified)
Physiological Parameters
AE (%) IR(g g 1 d ')
ke (d"1)
TTF
Study
52.0 0.42
0.155

Schlekat et al. 2004
Median Values and TTF



50.0 0.42
0.155
1.41

Blackworm (Lumbriculus variegatus)
Physiological Parameters
AE (%) IR(g g 1 d ')
ke (d"1)
TTF
Study

0.009

Riedel and Cole 2001

0.006

Riedel and Cole 2001
24.0 0.067
0.013

Riedel and Cole 2001
9.0 0.067
0.009

Riedel and Cole 2001
Median Values and TTF



16.5 0.067
0.0086
1.29

B-65

-------
Mayfly (Centroptilum triangulifer)a
Physiological Parameters


AE (%)
IR(g g1 d"1)
ke (d"1)
TTF Study
38.0
0.72
0.25
Riedel and Cole 2001
40.0
0.72
0.19
Riedel and Cole 2001
Median Values and TTF


39.0
0.72
0.22
1.28
a - not used because field TTF data available
3.2.2 Vertebrates
Bluegill (Lepomis macrochirus)a
Physiological Parameters
AE (%)
IRfeg-1*1) Md1)
TTF Study

34.0

Besser et al.
1993
22.0

Besser et al.
1993
24.0

Besser et al.
1993
36.0

Besser et al.
1993
30.0

Besser et al.
1993
32.0

Besser et al.
1993
43.0

Besser et al.
1993
40.0

Besser et al.
1993
37.0
0.041
Besser et al.
1993

0.031
Besser et al.
1993

0.034
Besser et al.
1993
36.0
0.031
Besser et al.
1993

0.038
Besser et al.
1993

0.038
Besser et al.
1993

0.008
Whitledge and Haywood

0.042
Whitledge and Haywood
Median Values and TTF
35.0	0.025 0.036 1.156s
a Not used because of availability of acceptable field-based TTF data
B-66

-------
Fathead Minnow (Pimephales promelas)
Physiological Parameters



AE (%) IR(g g1 d ')
ke (d"1)
TTF
Study
50.0


Presser and Luoma 2010
0.050


Bertram and Brooks 1986

0.029

Bertram and Brooks 1986

0.019

Bertram and Brooks 1986

0.3

Bertram and Brooks 1986

0.014

Bertram and Brooks 1986

0.013

Bertram and Brooks 1986

0.016

Bertram and Brooks 1986

0.012

Bertram and Brooks 1986

0.026

Bertram and Brooks 1986

0.018

Bertram and Brooks 1986

0.025

Bertram and Brooks 1986
Median Values and TTF



50.0 0.050
0.0185
1.35

Striped Bass (Morone saxatilis)
Physiological Parameters



AE (%) IR(g g1 d ')
ke (d"1)
TTF
Study
33 0.17
0.09

Baines et al. 2002
42 0.5
0.08

Baines et al. 2002
0.12


Buckel and Stoner 2004
0.16


Buckel and Stoner 2004
0.11


Buckel and Stoner 2004
0.08


Buckel and Stoner 2004
Median Values and TTF



37.5 0.335
0.085
1.48

TTF calculated from only Baines et al. (2002) because it had complete data.
B-67

-------
3.3 TTF values from field data
3.3.1 Invertebrates
Caig	=	Selenium concentration in algae (mg/kg)
Cdet	=	Selenium concentration in detritus (mg/kg)
Csed	=	Selenium concentration in sediment (mg/kg)
Cmvert	=	Selenium concentration in invertebrate tissue (mg/kg)
Cpart	=	Average selenium concentration in particulate material ^Cal3+C<^t+Csed
Ratio	=
Cpart
Scuds (Amphipoda)
Study

Site
r
ale
Cdet
r
v sed
r
v n;trt
c
^invert
Ratio
Birkner 1978
29
8.80

15.40
12.10
18.40
1.52
Birkner 1978
20
3.00

41.00
22.00
11.40
0.52
Birkner 1978
7
0.18

2.80
1.49
2.90
1.95
Birkner 1978
19
16.80

1.20
9.00
4.30
0.48
Birkner 1978
30
17.30

47.30
32.30
22.50
0.70
Birkner 1978
3
0.10

0.30
0.20
2.30
11.50
Birkner 1978
22
4.60

44.00
24.30
7.60
0.31
Birkner 1978
23
7.80

10.80
9.30
11.30
1.22
Lambing et al. 1994
S46
2.30


2.30
3.20
1.39
Saiki et al.
1993
ET6
1.03
1.15

1.09
0.44
0.40
Saiki et al.
1993
ET6
1.03
1.15

1.09
0.86
0.79
Saiki et al.
1993
GT5
4.50
14.95

9.73
4.60
0.47
Saiki et al.
1993
GT5
4.50
14.95

9.73
3.30
0.34
Saiki et al.
1993
GT4
1.39
8.40

4.90
3.40
0.69
Saiki et al.
1993
GT4
1.39
8.40

4.90
3.70
0.76
Saiki et al.
1993
SJR2
1.25
5.00

3.13
3.80
1.22
Saiki et al.
1993
SJR2
1.25
5.00

3.13
2.80
0.90
Saiki et al.
1993
SJR3
0.45
1.25

0.85
1.50
1.77
Saiki et al.
1993
SJR3
0.45
1.25

0.85
1.10
1.30
Saiki et al.
1993
SJR1
0.22
0.50

0.36
0.89
2.47
Saiki et al.
1993
SJR1
0.22
0.50

0.36
1.30
3.61
Saiki et al.
1993
ET7
0.16
0.76

0.46
1.10
2.42
Saiki et al.
1993
ET7
0.16
0.76

0.46
1.10
2.42
B-68

-------
Scuds (Amphipoda)
-Inverts
23 
20 -

15 -


o
10 

5 


0 
1

10	20
c
v-p*rtic.
O
Median ratio: 1.22
40
R
F
df
P
0.69
46.9
21
<0.001
Earthworms and Leeches (Annelida)
Study
Site
r
ale
^ (lei
f
v sc(l
c
v n;trt
c
^invert
Ratio
Lemly 1985
Badin Lake
8.20

0.91
4.56
8.10
1.78
Lemly 1985
Belews Lake
62.70

8.27
35.49
51.15
1.44
Lemly 1985
High Rock Lake
8.25

0.79
4.52
9.05
2.00
60
50
40
r _ 30
		
-inverts
20
10
0
10
,
20
Median ratio: 1.78
R
F
df
P
l
40
1.00
2426
1
<0.001
-pirtic.
B-69

-------
Midges (Chironomidae)
Study
Site
Calg
Cdet
Csed
C^art
Cjnvert
Ratio
Birkner 1978
29
8.80

15.40
12.10
58.20
4.81
Birkner 1978
19
16.80

1.20
9.00
15.30
1.70
Birkner 1978
30
17.30

47.30
32.30
59.30
1.84
Birkner 1978
3
0.10

0.30
0.20
2.50
12.50
Birkner 1978
22
4.60

44.00
24.30
18.80
0.77
Birkner 1978
27
10.35

6.50
8.43
26.70
3.17
Birkner 1978
12
2.30

0.30
1.30
7.70
5.92
Birkner 1978
23
7.80

10.80
9.30
34.20
3.68
Grasso et al. 1995
17
1.87

0.40
1.14
2.07
1.82
Lambing et al. 1994
S46
2.30


2.30
9.70
4.22
Saiki and Lowe 1987
Kesterson Pond 11
18.15
47.95
8.56
18.15
71.00
3.91
Saiki and Lowe 1987
Kesterson Pond 2
152.7
44.65
34.82
44.65
200.0
4.48
Saiki and Lowe 1987
Kesterson Pond 2
152.7
44.65
34.82
44.65
290.0
6.49
Saiki and Lowe 1987
Kesterson Pond 8
136.5
92.00
6.05
92.00
220.0
2.39
Saiki and Lowe 1987
San Luis Drain
67.00
275.0
79.90
79.90
190.0
2.38
Saiki and Lowe 1987
San Luis Drain
67.00
275.0
79.90
79.90
284.0
3.55
Saiki and Lowe 1987
Volta Pond 26
0.42
1.01
0.29
0.42
1.74
4.18
Saiki and Lowe 1987
Volta Pond 26
0.42
1.01
0.29
0.42
1.30
3.13
Saiki and Lowe 1987
Volta Pond 7

1.39
0.39
0.89
3.00
3.37
Saiki and Lowe 1987
Volta Pond 7

1.39
0.39
0.89
1.30
1.46
Saiki et al. 1993
ET6
1.03
1.15

1.09
0.58
0.53
Saiki et al. 1993
ET6
1.03
1.15

1.09
1.00
0.92
Saiki et al. 1993
GT5
4.50
14.95

9.73
8.90
0.92
Saiki et al. 1993
GT5
4.50
14.95

9.73
7.20
0.74
Saiki et al. 1993
GT4
1.39
8.40

4.90
5.40
1.10
Saiki et al. 1993
GT4
1.39
8.40

4.90
6.90
1.41
Saiki et al. 1993
SJR2
1.25
5.00

3.13
6.00
1.92
Saiki et al. 1993
SJR2
1.25
5.00

3.13
4.10
1.31
Saiki et al. 1993
SJR3
0.45
1.25

0.85
1.50
1.77
Saiki et al. 1993
SJR3
0.45
1.25

0.85
1.60
1.89
Saiki et al. 1993
SJR1
0.22
0.50

0.36
0.47
1.31
Saiki et al. 1993
SJR1
0.22
0.50

0.36
1.00
2.78
Saiki et al. 1993
ET7
0.16
0.76

0.46
0.53
1.16
Saiki et al. 1993
ET7
0.16
0.76

0.46
0.84
1.85
B-70

-------
Midges (Chironomidae)
inverts
partic.
100
Median ratio: 1.90
R
F
df
P
0.82
144.0
32
<0.001
Beetles (Coleoptera)
Study
Site
r
ale
^ (lei
r
v sed
r
v n;trt
r
^invert
Ratio
Schuler et al. 1990
Kesterson Pond 11
53.70

11.50
32.60
77.60
2.38
Schuleretal. 1990
Kesterson Pond 11
53.70

11.50
32.60
74.10
2.27
Schuler etal. 1990
Kesterson Pond 11
53.70

11.50
32.60
110.00
3.37
Schuleretal. 1990
Kesterson Pond 2
52.50

9.30
30.90
54.00
1.75
Schuleretal. 1990
Kesterson Pond 7
87.10

5.90
46.50
89.10
1.92
Schuleretal. 1990
Kesterson Pond 7
87.10

5.90
46.50
28.80
0.62
Schuleretal. 1990
Kesterson Pond 7
87.10

5.90
46.50
43.70
0.94
-"inverts
120 -I
100 
80 
60 
40 -
20
0
0
o
o
20
i
40
60
ratio:
1.92
R2:
0.20
F:
1.24
df:
5
P:
0.36
Not used because P > 0.05 and negative
slope.
'pirtic.
Water boatmen (Corixidae)
Study
Site
Calg
^ del Csed
Cpart
Cjnvert
Ratio
Birkner 1978
18
7.60
4.30
5.95
8.40
1.41
Birkner 1978
29
8.80
15.40
12.10
29.40
2.43
B-71

-------
Water boatmen (Corixidae)
Study
Site
Calg
Cdet
Csed
C^art
Cjnvert
Ratio
Birkner 1978
20
3.00

41.00
22.00
11.00
0.50
Birkner 1978
7
0.18

2.80
1.49
4.20
2.82
Birkner 1978
3
0.10

0.30
0.20
4.20
21.00
Birkner 1978
22
4.60

44.00
24.30
9.90
0.41
Birkner 1978
12
2.30

0.30
1.30
7.30
5.62
Birkner 1978
23
7.80

10.80
9.30
15.50
1.67
Lambing et al. 1994
S46
2.30


2.30
3.40
1.48
Rinellaetal. 1994
G
0.84

0.50
0.67
1.38
2.06
Rinellaetal. 1994
A
2.21

0.40
1.31
2.98
2.28
Rinellaetal. 1994
Q
1.42

0.50
0.96
2.00
2.08
Saiki and Lowe 1987
Kesterson Pond 11
18.15
47.95
8.56
18.15
24.00
1.32
Saiki and Lowe 1987
Kesterson Pond 11
18.15
47.95
8.56
18.15
16.00
0.88
Saiki and Lowe 1987
Kesterson Pond 8
136.50
92.00
6.05
92.00
20.00
0.22
Saiki and Lowe 1987
Kesterson Pond 8
136.50
92.00
6.05
92.00
24.00
0.26
Saiki and Lowe 1987
Volta Pond 26
0.42
1.01
0.29
0.42
2.15
5.17
Saiki and Lowe 1987
Volta Pond 26
0.42
1.01
0.29
0.42
0.87
2.10
Saiki and Lowe 1987
Volta Pond 7

1.39
0.39
0.89
1.76
1.98
Saiki and Lowe 1987
Volta Pond 7

1.39
0.39
0.89
1.53
1.72
Schuler etal. 1990
Kesterson Pond 11
53.70

11.50
32.60
15.90
0.49
Schuler etal. 1990
Kesterson Pond 11
53.70

11.50
32.60
64.60
1.98
Schuler etal. 1990
Kesterson Pond 11
53.70

11.50
32.60
15.10
0.46
Schuler etal. 1990
Kesterson Pond 2
52.50

9.30
30.90
20.00
0.65
Schuler etal. 1990
Kesterson Pond 2
52.50

9.30
30.90
10.00
0.32
Schuler etal. 1990
Kesterson Pond 7
87.10

5.90
46.50
23.00
0.49
Schuler etal. 1990
Kesterson Pond 7
87.10

5.90
46.50
30.90
0.66
Schuler etal. 1990
Kesterson Pond 7
87.10

5.90
46.50
6.46
0.14
Rinella and Schuler
18
0.59


0.59
2.70
4.58
1992







-"inverts
pirtic.
100
Median ratio:
1.48
R2:
0.25
F:
9.17
df:
27
P:
<0.001
B-72

-------
Crayfish (Astacidae)
Study

Site
Calg
Cdet
Csed
C^art
Cjnvert
Ratio
Birkner 1978
29
8.80

15.40
12.10
23.30
1.93
Birkner 1978
19
16.80

1.20
9.00
10.10
1.12
Birkner 1978
30
17.30

47.30
32.30
36.80
1.14
Birkner 1978
22
4.60

44.00
24.30
11.30
0.47
Birkner 1978
27
10.35

6.50
8.43
20.00
2.37
Butler et al
1993
SP2
1.60

0.50
1.05
2.60
2.48
Butler et al
1993
SP2
1.60

0.50
1.05
2.90
2.76
Butler et al
1995
AK
0.45

0.20
0.33
0.76
2.34
Butler et al
1995
AK
0.45

0.20
0.33
0.79
2.43
Butler et al
1995
DD
0.88

0.70
0.79
0.62
0.78
Butler et al
1995
DD
0.88

0.70
0.79
1.10
1.39
Butler et al
1995
HD1
0.59


0.59
0.86
1.46
Butler et al
1995
HD1
0.59


0.59
0.79
1.34
Butler et al
1995
HD2
0.45

0.20
0.32
0.96
2.98
Butler et al
1995
HD2
0.45

0.20
0.32
1.00
3.10
Butler et al
1995
ME2
1.11

1.10
1.10
1.10
1.00
Butler et al
1995
ME2
1.11

1.10
1.10
1.40
1.27
Butler et al
1995
ME4
1.04

0.50
0.77
1.30
1.69
Butler et al
1995
ME4
1.04

0.50
0.77
1.80
2.35
Butler et al
1995
ME3
0.82

0.40
0.61
1.40
2.30
Butler et al
1995
ME3
0.82

0.40
0.61
3.70
6.07
Butler et al
1995
NW
3.45

1.60
2.53
4.20
1.66
Butler et al
1995
NW
3.45

1.60
2.53
3.30
1.31
Butler et al
1995
SD
0.77

0.50
0.64
1.40
2.20
Butler et al
1995
SD
0.77

0.50
0.64
1.40
2.20
Butler et al
1995
YJ2
0.31

0.10
0.21
1.40
6.83
Butler et al
1995
YJ2
0.31

0.10
0.21
1.50
7.32
Butler et al
1997
CHK
1.19


1.19
0.90
0.76
Butler et al
1997
MN2
0.79


0.79
0.83
1.06
Butler et al
1997
MUD2
1.30


1.30
3.10
2.38
Butler et al
1997
MUD2
1.30


1.30
3.80
2.92
Butler et al
1997
TRH
1.25


1.25
0.98
0.78
Butler et al
1997
TRH
1.25


1.25
1.60
1.28
Saiki et al.
1993
ET6
1.03
1.15

1.09
0.67
0.62
Saiki et al.
1993
ET6
1.03
1.15

1.09
0.83
0.76
Saiki et al.
1993
GT5
4.50
14.95

9.73
5.20
0.53
Saiki et al.
1993
GT5
4.50
14.95

9.73
4.40
0.45
Saiki et al.
1993
GT4
1.39
8.40

4.90
3.10
0.63
Saiki et al.
1993
GT4
1.39
8.40

4.90
3.20
0.65
Saiki et al.
1993
SJR2
1.25
5.00

3.13
1.70
0.54
Saiki et al.
1993
SJR2
1.25
5.00

3.13
1.90
0.61
B-73

-------
Crayfish (Astacidae)
Study
Site
r
ale
^ del
C C
v sed v n;trt
r
^invert
Ratio
Saikietal. 1993
SJR3
0.45
1.25
0.85
0.77
0.91
Saiki et al. 1993
SJR3
0.45
1.25
0.85
1.30
1.53
Saikietal. 1993
SJR1
0.22
0.50
0.36
0.50
1.39
Saikietal. 1993
SJR1
0.22
0.50
0.36
0.74
2.06
Saikietal. 1993
ET7
0.16
0.76
0.46
0.87
1.91
Saikietal. 1993
ET7
0.16
0.76
0.46
0.85
1.87
-'inverts
40
35
30
25
20
15
10
5
0

10
20
^"partic.
30
40
Median ratio: 1.46
R
F
df
P
0.74
130.8
45
<0.001
True flies (Diptera)
Study
Site
Calg
^ del Csed
C^art
Cjnvert
Ratio
Schuler et al. 1990
Kesterson Pond 11
53.70
11.50
32.60
126.00
3.87
Schuler et al. 1990
Kesterson Pond 11
53.70
11.50
32.60
85.10
2.61
Schuler et al. 1990
Kesterson Pond 2
52.50
9.30
30.90
117.00
3.79
Schuler et al. 1990
Kesterson Pond 2
52.50
9.30
30.90
93.30
3.02
Schuler etal. 1990
Kesterson Pond 2
52.50
9.30
30.90
105.00
3.40
Schuler etal. 1990
Kesterson Pond 7
87.10
5.90
46.50
95.50
2.05
Schuler etal. 1990
Kesterson Pond 7
87.10
5.90
46.50
97.70
2.10
Schuler etal. 1990
Kesterson Pond 7
87.10
5.90
46.50
102.00
2.19
140 -i






120 
o
o

Median ratio:
2.81

100 
o-		
o





80 
o


R2:
0.07

c
inverts gQ .



F:
0.46




df:
6

40 



P:
0.65

20
0
Not used because P > 0.05 and negative
slope.
20	40
Cpartic.
60
B-74

-------
Mayflies (Ephemeroptera)
Study
Site
Calg
Cdet
Csed
C' [);nl
Cjnvert
Ratio
Rinellaetal. 1994
A
2.21

0.40
1.31
9.65
7.39
Casey 2005
Deerlick Creek

1.00
0.20
0.60
6.40
10.67
Casey 2005
Luscar Creek
5.50
3.20
2.40
3.20
8.20
2.56
Casey 2005
Deerlick Creek

1.00
0.20
0.60
5.70
9.50
Casey 2005
Luscar Creek
5.50
3.20
2.40
3.20
9.70
3.03
Casey 2005
Deerlick Creek

1.00
0.20
0.60
6.80
11.33
Casey 2005
Luscar Creek
5.50
3.20
2.40
3.20
12.30
3.84
Conley et al. 2009
Plate 10A
4.40


4.40
9.70
2.20
Conley et al. 2009
Plate 20A
25.50


25.50
34.80
1.36
Conley et al. 2009
Plate 20B
17.50


17.50
56.70
3.24
Conley et al. 2009
Plate 20C
8.70


8.70
16.20
1.86
Conley et al. 2009
Plate 20D
11.30


11.30
27.50
2.43
Conley et al. 2009
Plate 5A
2.20


2.20
4.20
1.91
Conley et al. 2009
Plate 5B
2.00


2.00
5.70
2.85
Conley etal. 2011
2x-High
40.90


40.90
37.30
0.91
Conley etal. 2011
2x-Low
9.50


9.50
14.10
1.48
Conley etal. 2011
2x-Medium
19.90


19.90
21.60
1.09
Conley et al. 2013
Control
2.20


2.20
5.10
2.32
Conley et al. 2013
Selenate-high
36.80


36.80
59.80
1.63
Conley et al. 2013
Selenate-low
12.80


12.80
31.70
2.48
Conley et al. 2013
Selenite-high
36.70


36.70
78.40
2.14
Conley et al. 2013
Selenite-low
12.80


12.80
29.80
2.33
-'inverts
90
80
70
60
50
40
30
20
10
0
10
20
c
30
40
Median ratio: 2.38
R2:
0.75
F:
59.19
df:
20
P:
<0.001
particulate
B-75

-------
Snails (Gastropoda)
Study
Site
r
ale
C C
^det ^sed
r
^oart
r
^invert
Ratio
Butler etal. 1995
WC
3.30
1.50
2.40
3.70
1.54
Butler etal. 1995
WC
3.30
1.50
2.40
3.90
1.63
Butler etal. 1995
WC
3.30
1.50
2.40
2.00
0.83
Butler etal. 1997
DCP1
1.00
2.10
1.55
3.50
2.26
Butler etal. 1997
MNP2
5.40
6.70
6.05
2.00
0.33
Butler etal. 1997
CHP
4.00
2.10
3.05
19.00
6.23
Butler etal. 1997
LCHP1
0.33
1.10
0.72
0.32
0.45
'Inverts
20
15
10
-o e
o
-jQ_
2	4
r
vprtic.
Median ratio:
1.54
R2:
0.01
F:
0.07
df:
5
P:
0.93
Not used because P >
O
O
Lh
Zooplankton
Study
Site
Calg
Cdet
Csed
C^art
Cjnvert
Ratio
Birkner 1978
29
8.80

15.40
12.10
31.30
2.59
Birkner 1978
20
3.00

41.00
22.00
11.00
0.50
Birkner 1978
7
0.18

2.80
1.49
3.30
2.22
Birkner 1978
19
16.80

1.20
9.00
7.70
0.86
Birkner 1978
3
0.10

0.30
0.20
3.40
17.00
Birkner 1978
27
10.35

6.50
8.43
42.50
5.04
Birkner 1978
12
2.30

0.30
1.30
5.80
4.46
Birkner 1978
23
7.80

10.80
9.30
15.40
1.66
Bowie et al. 1996
Hyco Reservoir
27.0


27.0
23.0
0.85
Lambing et al. 1988
12
1.40

0.30
0.85
2.60
3.06
Saiki and Lowe 1987
Kesterson Pond 11
18.15
47.95
8.56
18.15
68.30
3.76
Saiki and Lowe 1987
Kesterson Pond 2
152.70
44.65
34.82
44.65
83.00
1.86
Saiki and Lowe 1987
Kesterson Pond 8
136.50
92.00
6.05
92.00
100.00
1.09
Saiki and Lowe 1987
Volta Pond 26
0.42
1.01
0.29
0.42
1.46
3.51
Saiki and Lowe 1987
Volta Pond 7

1.39
0.39
0.89
2.90
3.26
Saiki and Lowe 1987
Volta Wasteway
0.87
2.03
0.24
0.87
2.80
3.21
Saiki etal. 1993
ET6
1.03
1.15

1.09
1.20
1.10
Saiki etal. 1993
ET6
1.03
1.15

1.09
1.50
1.38
B-76

-------
Zooplankton
Study
Site
Calg
^ (lei Csed CDart
Cjnvert
Ratio
Saikietal. 1993
GT5
4.50
14.95
9.73
2.40
0.25
Saikietal. 1993
GT5
4.50
14.95
9.73
5.40
0.56
Saikietal. 1993
GT4
1.39
8.40
4.90
4.50
0.92
Saikietal. 1993
GT4
1.39
8.40
4.90
4.40
0.90
Saikietal. 1993
SJR2
1.25
5.00
3.13
2.60
0.83
Saikietal. 1993
SJR2
1.25
5.00
3.13
4.30
1.38
Saikietal. 1993
SJR3
0.45
1.25
0.85
1.60
1.89
Saikietal. 1993
SJR3
0.45
1.25
0.85
1.80
2.12
Saikietal. 1993
SJR1
0.22
0.50
0.36
1.40
3.89
Saikietal. 1993
SJR1
0.22
0.50
0.36
1.30
3.61
Saikietal. 1993
ET7
0.16
0.76
0.46
0.63
1.38
Saikietal. 1993
ET7
0.16
0.76
0.46
1.40
3.08
140
120
100 
80
60
40
20
0
jzf
20
40
c,
60

100
Median ratio: 1.89
R
F
df
P
0.71
76.3
31
<0.001
particulate
B-77

-------
Special case of Odonates (Damsclflics and Dragonflics) consuming invertebrates
n	= Number of invertebrate food species co-occurring with an Odonate
species.
CPart	= Average selenium concentration in particulate material
(mg/kg): (c-?tcy+c"-)
Cfood	= Median selenium concentration in all invertebrate tissues that co-
occur with an Odonate species (mg/kg)
Cdamsel	= Selenium concentration in damselfly tissue (mg/kg)
Cdragon	= Selenium concentration in dragonfly tissue (mg/kg)
Ratio	= Cfod cdamsel Qr cdragon
Cpart	Cfood	Cfood
Co-occurring potential food species of damselflies and dragonflies (Odonata)
Study
Site
Co-occurs with:
n
r
^oart
Cfood
Ratio
Saiki and Lowe 1987
Kesterson
Pond 11
dragonflies
4
18.15
47.5
2.62
Saiki and Lowe 1987
Kesterson
Pond 2
dragonflies
4
44.65
206.5
4.62
Saiki and Lowe 1987
Kesterson
Pond 2
dragonflies
4
44.65
206.5
4.62
Saiki and Lowe 1987
Kesterson
Pond 8
dragonflies
5
92.00
120
1.30
Saiki and Lowe 1987
Kesterson
Pond 8
dragonflies
5
92.00
120
1.30
Saiki and Lowe 1987
Volta Pond 26
dragonflies
4
0.42
1.52
3.65
Saiki and Lowe 1987
Volta Pond 26
dragonflies
4
0.42
1.52
3.65
Saiki and Lowe 1987
Volta Pond 7
dragonflies
5
0.89
1.53
1.72
Saiki and Lowe 1987
Volta Pond 7
dragonflies
5
0.89
1.53
1.72
Saiki and Lowe 1987
Volta
Wasteway
dragonflies
2
0.87
1.83
2.10
Schuler et al. 1990
Kesterson
Pond 11
dragonflies
10
32.60
75.85
2.33
Schuler et al. 1990
Kesterson
Pond 11
dragonflies
10
32.60
75.85
2.33
Schuler et al. 1990
Kesterson
Pond 2
dragonflies
8
30.90
93.3
3.02
Schuler etal. 1990
Kesterson
Pond 2
dragonflies
8
30.90
93.3
3.02
B-78

-------
Co-occurring potential food species of damselflies and dragonflies (Odonata)
Study

Site
Co-occurs with:
n
C^art
Cfood
Ratio
Schuler et al.
1990
Kesterson
dragonflies
11
46.50
69.2
1.49


Pond 7





Schuler et al.
1990
Kesterson
dragonflies
11
46.50
69.2
1.49


Pond 7





Birkner 1978

29
damselflies
3
12.10
29.4
2.43
Birkner 1978

20
damselflies
2
22.00
11.2
0.51
Birkner 1978

7
damselflies
2
1.49
3.55
2.39
Birkner 1978

19
damselflies
2
9.00
9.8
1.09
Birkner 1978

30
damselflies
2
32.30
40.9
1.27
Birkner 1978

3
damselflies
3
0.20
2.5
12.50
Birkner 1978

22
damselflies
3
24.30
9.9
0.41
Birkner 1978

27
damselflies
1
8.43
26.7
3.17
Birkner 1978

23
damselflies
3
9.30
15.5
1.67
Grassoetal. 1995
17
damselflies
1
1.14
2.07
1.82
250 -I







200 

o





150 



Median
ratio:
2.21

Cfood |0O 


o

R2:
0.54

50 -

o/-o


F:
28.7

Q0"
9 " o


df:
24

0 *
t 
Oo
	1	1	1	
	i	i

P:
<0.001

0 20 40 60 80 100
r
'-particulate
Damselflies (Anisoptera)
Study
Site
Cfood
r
^damsel
Ratio
Birkner 1978
29
29.4
55
1.87
Birkner 1978
4
1.95
1.8
0.92
Birkner 1978
25
18.7
21.9
1.17
Birkner 1978
20
11.2
18.7
1.67
Birkner 1978
7
3.55
4.4
1.24
Birkner 1978
19
9.8
28.4
2.90
Birkner 1978
6
4.2
11.1
2.64
Birkner 1978
30
40.9
53.3
1.30
Birkner 1978
3
2.5
3.1
1.24
Birkner 1978
22
9.9
15.8
1.60
B-79

-------
Damselflies (Anisoptera)
Study

Site
Cfood
c
^damsel
Ratio
Birkner 1978

27
26.7
45.1
1.69
Birkner 1978

23
15.5
18.4
1.19
Birkner 1978

11
5.9
7.7
1.31
Grassoetal. 1995
17
2.07
1.75
0.85
Grassoetal. 1995
9
8.2
6.98
0.85
100 -I





80 -


Median ratio:
1.30x2.21

60 -
r
damsel
40 
O
o

(damselfly food to particulate) = 2.8
R2: 0.89
F: 104.4
8
20 


df:
13




P:
<0.001

20
40
'food
Dragonflies (Zygoptera)
Study
Site
Cfood
Cdragon
Ratio
Mason et al. 2000
BK
1.845
1.665
0.90
Mason et al. 2000
HCRT
4.305
2.81
0.65
Saiki and Lowe 1987
Kesterson Pond 11
47.5
53
1.12
Saiki and Lowe 1987
Kesterson Pond 2
206.5
155
0.75
Saiki and Lowe 1987
Kesterson Pond 2
206.5
171
0.83
Saiki and Lowe 1987
Kesterson Pond 8
120
95.5
0.80
Saiki and Lowe 1987
Kesterson Pond 8
120
105
0.88
Saiki and Lowe 1987
Volta Pond 26
1.52
1.4
0.92
Saiki and Lowe 1987
Volta Pond 26
1.52
1.42
0.93
Saiki and Lowe 1987
Volta Pond 7
1.53
1.2
0.78
Saiki and Lowe 1987
Volta Pond 7
1.53
1.4
0.92
Saiki and Lowe 1987
Volta Wasteway
1.83
2.5
1.37
Schuleretal. 1990
Kesterson Pond 11
75.85
63.1
0.83
Schuleretal. 1990
Kesterson Pond 11
75.85
95.5
1.26
Schuleretal. 1990
Kesterson Pond 2
93.3
110
1.18
Schuleretal. 1990
Kesterson Pond 2
93.3
65
0.70
Schuleretal. 1990
Kesterson Pond 7
69.2
61.7
0.89
Schuleretal. 1990
Kesterson Pond 7
69.2
56.2
0.81
Sorenson & Schwarzbach 1991
5
0.42
0.49
1.17
B-80

-------
Dragonflies (Zygoptera)
200
150
-"dragon
4^-

Median ratio: 0.89x2.21
(damselfly food to particulate) = 1.97
R2:
0.95
F:
343.5
df:
17
P:
<0.001
100	200
Cfood
300
3.3.2 Vertebrates
Cinvert = Selenium concentration in invertebrate tissue (jj.g/g)
Cfish	= Average selenium concentration in the whole-body of fish (jj.g/g)
Ratio =
C invert
Black bullhead (Ameiurus melas)
Study
Site
Cjnvert
Cfish
Ratio
Butler etal. 1991
7
29.80
39.00
1.31
GEI2013
SWA1
2.81
2.37
0.84
GEI2013
SWA1
2.81
2.73
0.97
GEI 2013
SWA1
2.81
3.96
1.41
GEI 2013
SWA1
2.81
1.95
0.70
GEI 2013
SWA1
2.81
3.21
1.14
Mueller etal. 1991
R2
6.40
9.70
1.52
Mueller etal. 1991
R2
6.40
9.20
1.44
Mueller etal. 1991
R1
8.70
7.40
0.85
Lemly 1985
Badin Lake
5.70
2.58
0.45
Lemly 1985
Belews Lake
51.15
17.32
0.34
Lemly 1985
High Rock Lake
9.05
3.24
0.36
GEI 2014
SC-6
27.54
8.42
0.31
B-81

-------
Black bullhead (Ameiurus melas)
45
40
35
30
25
20
15
10
5
0
I O O

Median ratio:
0.85

R2:
0.44
o
F:
8.52

df:
11

P:
0.006
10	20	30	40
^-Invvrit
50
60
Black crappie (Pomoxis nigromaculatus)
Study
Site
Cjnvert
Cfish
Ratio
Butler etal. 1995
Totten Reservoir
1.07
2.50
2.35
Butler etal. 1995
Summit Reservoir
1.85
1.70
0.92
Peterson et al. 1991
Ocean Lake, west side
3.83
4.20
1.10
Peterson et al. 1991
Ocean Lake, west side
3.83
6.32
1.65
Mueller etal. 1991
Lake Meredith near
6.40
13.00
2.03

Ordway, CO



Lambing et al. 1994
Priest Butte Lakes near
14.00
39.00
2.79

Choteau



Lambing et al. 1994
Priest Butte Lakes near
14.00
41.00
2.93

Choteau



Lambing et al. 1994
Priest Butte Lakes near
14.00
47.00
3.36

Choteau



Lambing et al. 1994
Priest Butte Lakes near
15.00
40.00
2.67

Choteau



Lambing et al. 1994
Priest Butte Lakes near
15.00
57.00
3.80

Choteau



Lambing et al. 1994
Priest Butte Lakes near
15.00
63.00
4.20

Choteau



B-82

-------
Black crappie (Pomoxis nigromaculatus)
Median ratio
Blacknose dace (Rhinichthys atratulus)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
Cabin Creek, C-CC1
4.40
4.65
1.06
GEI2014
Cabin Creek, C-CC1
4.40
4.74
1.08
GEI 2014
Cabin Creek, C-CC1
4.40
4.95
1.12
GEI 2014
Cabin Creek, C-CC1
4.40
4.69
1.06
GEI 2014
Cabin Creek, C-CC1
4.40
3.98
0.90
GEI 2014
Cabin Creek, C-CC2
5.56
3.46
0.62
GEI 2014
Cabin Creek, C-CC2
5.56
3.38
0.61
GEI 2014
Cabin Creek, C-CC2
5.56
3.95
0.71
GEI 2014
Cabin Creek, C-CC2
5.56
4.36
0.78
GEI 2014
Cabin Creek, C-CC2
5.56
4.39
0.79
GEI 2014
Coal Fork, C-CF1
3.39
3.58
1.06
GEI 2014
Coal Fork, C-CF 1
3.39
3.09
0.91
GEI 2014
Coal Fork, C-CF 1
3.39
3.37
0.99
GEI 2014
Coal Fork, C-CF 1
3.39
2.64
0.78
GEI 2014
Coal Fork, C-CF 1
3.39
3.42
1.01
GEI 2014
Hazy Creek, C-HC1
8.03
3.99
0.50
GEI 2014
Hazy Creek, C-HC1
8.03
5.88
0.73
GEI 2014
Hazy Creek, C-HC1
8.03
4.46
0.56
GEI 2014
Hazy Creek, C-HC1
8.03
6.55
0.82
GEI 2014
Hazy Creek, C-HC1
8.03
3.98
0.50
GEI 2014
Laurel Fork, C-LF1
12.73
5.36
0.42
GEI 2014
Laurel Fork, C-LF1
12.73
7.99
0.63
GEI 2014
Laurel Fork, C-LF1
12.73
8.72
0.68
GEI 2014
Laurel Fork, C-LF1
12.73
5.49
0.43
GEI 2014
Tenmile Fork, C-TF1
20.00
7.62
0.38
GEI 2014
Tenmile Fork, C-TF1
20.00
10.56
0.53
B-83

-------
Blacknose dace (Rhinichthys atratulus)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
Tenmile Fork, C-TF1
20.00
8.02
0.40
GEI2014
Tenmile Fork, C-TF1
20.00
5.63
0.28
GEI 2014
Tenmile Fork, C-TF1
Jack Smith (Bear)
20.00
5.68
0.28
GEI 2014
Branch, H-JSB1
Jack Smith (Bear)
4.03
2.81
0.70
GEI 2014
Branch, H-JSB1
Jack Smith (Bear)
4.03
1.86
0.46
GEI 2014
Branch, H-JSB1
Jack Smith (Bear)
4.03
1.78
0.44
GEI 2014
Branch, H-JSB1
Jack Smith (Bear)
4.03
2.47
0.61
GEI 2014
Branch, H-JSB1
4.03
2.55
0.63
GEI 2014
Lukey Fork, H-LF1
9.09
5.32
0.59
GEI 2014
Mud River, H-MR3
3.86
8.72
2.26
GEI 2014
Mud River, H-MR6
2.49
3.80
1.53
GEI 2014
Mud River, H-MR6
Sugartree Branch, H-
2.49
2.93
1.18
GEI 2014
SB1
Sugartree Branch, H-
10.62
9.82
0.92
GEI 2014
SB1
Sugartree Branch, H-
10.62
7.29
0.69
GEI 2014
SB1
Sugartree Branch, H-
10.62
11.14
1.05
GEI 2014
SB1
Sugartree Branch, H-
10.62
4.85
0.46
GEI 2014
SB1
10.62
7.16
0.67
GEI 2014
Stanley Fork, H-SF1
21.05
18.21
0.87
20
18
16
14
12
rO"
O
o
o.
o
10
JL
o
o



Median ratio:
0.71
S--


8
R2:
0.52
o
F:
48.97

df:
45

P:
<0.001
20
25
C,.
B-84

-------
Bluegill (Lepomis macrochirus)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1995
XT
1.07
2.30
2.16
Hermanutz et al. 1996
MSO II
16.63
24.29
1.46
Hermanutz et al. 1996
MSO III
5.55
13.77
2.48
Hermanutz et al. 1996
MSO I
21.19
18.28
0.86
Hermanutz et al. 1996
MSO I
21.19
18.13
0.86
Hermanutz et al. 1996
MSO II
17.30
20.99
1.21
Hermanutz et al. 1996
MSO II
5.05
4.88
0.97
Hermanutz et al. 1996
MSO I
0.87
1.55
1.78
Hermanutz et al. 1996
MSO II
1.70
1.55
0.91
Hermanutz et al. 1996
MSO III
1.20
1.83
1.52
Hermanutz et al. 1996
MSO III
10.00
10.32
1.03
Hermanutz et al. 1996
MSO III
3.95
4.21
1.06
Hermanutz et al. 1996
MSO II
17.30
16.76
0.97
Hermanutz et al. 1996
MSO II
5.05
3.86
0.76
Mueller etal. 1991
R1
8.70
5.20
0.60
Saikietal. 1993
ET6
0.85
2.20
2.60
Saikietal. 1993
ET6
0.85
1.40
1.66
Saikietal. 1993
GT5
4.90
6.40
1.31
Saikietal. 1993
GT5
4.90
5.00
1.02
Saikietal. 1993
GT4
4.05
4.50
1.11
Saikietal. 1993
GT4
4.05
4.30
1.06
Saikietal. 1993
SJR2
3.30
3.30
1.00
Saikietal. 1993
SJR2
3.30
2.70
0.82
Saikietal. 1993
SJR3
1.50
2.00
1.33
Saikietal. 1993
SJR3
1.50
1.90
1.27
Saikietal. 1993
SJR1
0.95
0.87
0.92
Saikietal. 1993
SJR1
0.95
1.40
1.48
Saikietal. 1993
ET7
0.86
1.20
1.40
Saikietal. 1993
ET7
0.86
1.20
1.40
Crutchfield 2000
transect 3
21.80
19.91
0.91
Crutchfield 2000
transect 3
21.80
16.72
0.77
Crutchfield 2000
transect 3
17.90
19.91
1.11
Crutchfield 2000
transect 3
20.70
16.26
0.79
Crutchfield 2000
transect 3
20.35
29.87
1.47
Crutchfield 2000
transect 3
23.40
27.59
1.18
Crutchfield 2000
transect 3
15.20
23.10
1.52
Crutchfield 2000
transect 3
16.95
28.96
1.71
Crutchfield 2000
transect 3
16.95
19.91
1.17
Crutchfield 2000
transect 3
11.95
12.69
1.06
Crutchfield 2000
transect 3
11.40
18.09
1.59
Crutchfield 2000
transect 3
9.25
4.56
0.49
Crutchfield 2000
transect 3
9.25
5.40
0.58
B-85

-------
Bluegill (Lepomis macrochirus)
Study

Site
r
^invert

Cfish
Ratio
Crutchfield 2000

transect 3
8.60

4.56
0.53
Crutchfield 2000

transect 3
8.60

4.56
0.53
Crutchfield 2000

transect 4
30.70

51.60
1.68
Crutchfield 2000

transect 4
30.00

30.78
1.03
Crutchfield 2000

transect 4
33.20

31.69
0.95
Crutchfield 2000

transect 4
48.90

37.09
0.76
Crutchfield 2000

transect 4
38.55

49.78
1.29
Crutchfield 2000

transect 4
49.30

43.40
0.88
Crutchfield 2000

transect 4
43.90

22.65
0.52
Crutchfield 2000

transect 4
33.25

32.60
0.98
Crutchfield 2000

transect 4
25.40

18.09
0.71
Crutchfield 2000

transect 4
20.90

16.26
0.78
Crutchfield 2000

transect 4
20.90

26.22
1.25
Crutchfield 2000

transect 4
15.70

12.69
0.81
Crutchfield 2000

transect 4
15.70

9.04
0.58
Crutchfield 2000

transect 4
16.45

8.13
0.49
Crutchfield 2000

transect 4
18.25

9.96
0.55
Bowie et al. 1996

Hyco Reservoir
40.00

41.00
1.03
*0
AQ -

 o
oy'
Median ratio:
1.03

:u 0
O Q
10  .0
V
o 
14
o

-------
Bluehead sucker (Catostomus discobolus)
Study

Site
c
^invert
Cfish
Ratio
Butler et al.
1993
D1
1.20
2.80
2.33
Butler et al.
1993
B1
1.25
1.90
1.52
Butler et al.
1993
B1
1.25
2.20
1.76
Butler et al.
1995
ME2
1.25
0.83
0.66
Butler et al.
1995
ME2
1.25
1.30
1.04
Butler et al.
1993
B2
1.35
1.80
1.33
Butler et al.
1995
SD
1.40
1.50
1.07
Butler et al.
1995
SD
1.40
1.80
1.29
Butler et al.
1993
D2
1.45
1.60
1.10
Butler et al.
1993
D2
1.45
2.30
1.59
Butler et al.
1993
PI
1.50
2.20
1.47
Butler et al.
1994
C0L1
1.50
1.60
1.07
Butler et al.
1994
RB3
1.60
13.00
8.13
Butler et al.
1995
YJ2
1.65
0.96
0.58
Butler et al.
1995
YJ2
1.65
2.80
1.70
Butler et al.
1994
NFK3
2.00
1.40
0.70
Butler et al.
1997
MN2
2.20
1.20
0.55
Butler et al.
1997
MUD
2.30
1.80
0.78
Butler et al.
1997
MUD
2.30
2.30
1.00
Butler et al.
1997
CHK
2.40
1.20
0.50
Butler et al.
1997
CHK
2.40
1.60
0.67
Butler et al.
1993
U1
2.45
4.80
1.96
Butler et al.
1995
SJ1
2.50
0.94
0.38
Butler et al.
1995
SJ1
2.50
1.20
0.48
Butler et al.
1995
SJ1
2.50
1.20
0.48
Butler et al.
1995
ME3
2.55
1.70
0.67
Butler et al.
1995
ME3
2.55
1.80
0.71
Butler et al.
1997
MN3
2.70
1.50
0.56
Butler et al.
1997
MN1
2.90
1.40
0.48
Butler et al.
1993
SP1
2.95
5.10
1.73
Butler et al.
1993
SP2
3.40
7.10
2.09
Butler et al.
1997
MUD2
3.45
2.50
0.72
Butler et al.
1997
MUD2
3.45
5.20
1.51
Butler et al.
1997
MUD2
3.45
5.60
1.62
Butler et al.
1993
F2
3.90
10.00
2.56
Butler et al.
1991
4
3.90
1.80
0.46
Butler et al.
1993
F2
4.80
0.94
0.20
Butler et al.
1994
BSW1
5.00
33.00
6.60
Butler et al.
1997
WBR
5.05
1.80
0.36
Butler et al.
1997
WBR
5.05
2.80
0.55
Butler et al.
1995
NW
5.10
7.20
1.41
Butler et al.
1995
NW
5.10
9.30
1.82
B-87

-------
Bluehead sucker (Catostomus discobolus)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1994
LZA1
19.00
9.00
0.47
Butler etal. 1994
RBI
21.00
22.00
1.05
Butler etal. 1994
GUN2
28.00
3.60
0.13
Median ratio
Brook stickleback (Culaea
inconstans)



Study
Site
Cjnvert
Cfish
Ratio
GEI2013
SWA1
2.81
4.40
1.57
GEI2013
SWA1
2.81
4.59
1.64
GEI 2013
SWA1
2.81
4.66
1.66
GEI 2013
SWA1
2.81
5.00
1.78
GEI 2013
SWA1
2.81
5.21
1.86
GEI 2013
SWA1
3.64
3.69
1.02
GEI 2013
SWA1
3.64
4.16
1.14
GEI 2013
SWA1
3.64
4.21
1.16
GEI 2013
SWA1
3.64
4.62
1.27
GEI 2013
SWA1
3.64
4.78
1.31
GEI 2013
SWA1
3.64
4.98
1.37
GEI 2013
SWA1
3.64
5.06
1.39
GEI 2013
SWA1
3.64
6.28
1.73
Lambing et al. 1994
S38
4.70
17.00
3.62
Lambing et al. 1994
S37
5.30
6.10
1.15
Lambing et al. 1994
S36
6.30
5.30
0.84
GEI 2014
Dry Creek, DC-2
3
3.14
0.92
GEI 2014
Dry Creek, DC-2
3
4.03
1.18
GEI 2014
Dry Creek, DC-2
3
3.76
1.10
GEI 2014
Dry Creek, DC-2
3
5.31
1.55
GEI 2014
Dry Creek, DC-2
3
4.59
1.34
GEI 2014
Dry Creek, DC-3
7
28.89
4.02
GEI 2014
Dry Creek, DC-3
7
66.07
9.20
B-88

-------
Brook stickleback (Culaea inconstans)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
Dry Creek, DC-3
7
24.43
3.40
GEI2014
Dry Creek, DC-3
7
25.36
3.53
GEI 2014
Dry Creek, DC-3
7
40.17
5.59
GEI 2014
Dry Creek, DC-3
9
25.80
2.80
GEI 2014
Dry Creek, DC-3
9
24.14
2.62
GEI 2014
Dry Creek, DC-3
9
22.46
2.43
GEI 2014
Dry Creek, DC-3
9
19.86
2.15
GEI 2013
SW2-1
6.60
21.14
3.21
GEI 2013
SW2-1
6.60
23.21
3.52
GEI 2013
SW2-1
6.60
23.64
3.58
GEI 2013
SW2-1
6.60
25.89
3.93
GEI 2013
SW2-1
6.60
27.71
4.20
GEI 2013
SW2-1
6.60
32.97
5.00
GEI 2013
SW2-1
6.60
34.54
5.24
GEI 2013
SW2-1
6.60
37.05
5.62
GEI 2013
SW2-1
6.60
39.26
5.95
GEI 2013
SW2-1
6.60
43.38
6.58
GEI 2013
SWB
7.06
15.74
2.23
GEI 2013
SWB
7.06
17.15
2.43
GEI 2013
SW1
7.82
9.96
1.27
GEI 2013
SW1
7.82
10.38
1.33
GEI 2013
SW1
7.82
10.58
1.35
GEI 2013
SW1
7.82
11.98
1.53
GEI 2013
SW11
8.41
6.36
0.76
GEI 2013
SW11
8.41
6.45
0.77
GEI 2013
SW2-1
9.14
21.09
2.31
Lambing et al. 1994
S34
14.00
35.00
2.50
Lambing et al. 1994
Sll
14.50
22.00
1.52
Lambing et al. 1994
Sll
14.50
26.00
1.79
70
60
50
40
30
20
10
0
i
Og  o o
o
o
o
10
^inverts
15
20
Median ratio: 1.79
R
F
df
P
0.27
18.48
50
<0.001
B-89

-------
Brook trout (Salvelinus fontinalis)
Study

Site
c
^invert
Cfish
Ratio
Hamilton and Buhl 2004
use
0.50
2.40
4.80
Mason et al. 2000
BK
1.43
1.21
0.84
Mason et al. 2000
BK
1.43
1.57
1.10
Mason et al. 2000
BK
1.43
1.90
1.33
Mason et al. 2000
HCRT
2.81
0.99
0.35
Mason et al. 2000
HCRT
2.81
1.59
0.57
Mason et al. 2000
HCRT
2.81
2.95
1.05
Butler etal. 1997
MN1
2.90
2.20
0.76
Hamilton and Buhl 2005
LGC
7.80
6.90
0.88
Hamilton and Buhl 2005
UGC
9.30
9.80
1.05
Hamilton and Buhl 2004
DVC
12.80
8.00
0.63
Hamilton and Buhl 2004
use
0.50
2.40
4.80
12 -I





10 

 /



8 
Cfish 6
4 

o
Median ratio:
R2:
0.88
0.83


O

F:
43.6

z
A


df:
9

5
c
10
15
invert
B-90

-------
Brown bullhead (Ameiurus nebulosus)
Study


Site

c
^invert
Cfish
Ratio
Rinellaand Schuler 1992


1.20
1.90
1.58
Mason et al.
2000

HCRT

2.81
0.22
0.08
Mason et al.
2000

HCRT

2.81
1.23
0.44
Mason et al.
2000

HCRT

2.81
1.83
0.65
2 1
2 -



o
"\o
o
Median ratio:
0.55

Cfisi 1
1 -
A



R2:
F:
df:
P
0.27
0.73
2
0 58

U
)
1 2
^"invert.
3
Not used because P > 0.05 and negative
slope
Brown trout (Salmo trutta)
Study


Site

Cjnvert
Cfish
Ratio
Butler et al.
1991

10

4.80
2.00
0.42
Butler et al.
1991

12

2.80
5.40
1.93
Butler et al.
1991

4

3.90
3.30
0.85
Butler et al.
1991

4

3.90
3.50
0.90
Butler et al.
1991

3

6.20
3.50
0.56
Butler et al.
1993

SP2

3.40
3.40
1.00
Butler et al.
1993

SP2

2.75
1.20
0.44
Butler et al.
1993

B2

1.35
2.40
1.78
Butler et al.
1993

B2

1.35
2.70
2.00
Butler et al.
1993

B2

1.35
2.70
2.00
Butler et al.
1993

B1

1.25
4.20
3.36
Butler et al.
1993

D2

1.45
3.50
2.41
Butler et al.
1993

D2

1.45
3.50
2.41
Butler et al.
1993

D2

1.45
3.20
2.21
Butler et al.
1993

PI

1.95
3.30
1.69
Butler et al.
1993

LP2

1.00
1.70
1.70
Butler et al.
1993

LP2

1.00
2.10
2.10
Butler et al.
1993

LP2

1.00
1.60
1.60
Butler et al.
1993

LP3

1.12
2.10
1.88
Butler et al.
1993

LP3

1.12
2.80
2.51
Butler et al.
1993

LP4

3.20
1.80
0.56
B-91

-------
Brown trout (Salmo trutta)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1993
R2
3.90
5.40
1.38
Butler etal. 1993
R2
3.90
6.70
1.72
Butler etal. 1993
R2
3.70
5.90
1.59
Butler etal. 1993
ST2
4.10
6.00
1.46
Butler etal. 1994
GUN2
28.00
49.45
1.77
Butler etal. 1994
GUN2
28.00
5.90
0.21
Butler etal. 1994
HCC1
21.00
21.98
1.05
Butler etal. 1994
HCC1
21.00
42.00
2.00
Butler etal. 1994
NFK3
2.00
5.00
2.50
Butler etal. 1994
SMF
4.80
21.44
4.47
Butler etal. 1994
SMF
4.80
5.26
1.10
Butler etal. 1994
SMF
4.80
8.40
1.75
Butler etal. 1994
SMF
4.80
9.40
1.96
Formation 2012
CC-1A
12.24
10.51
0.86
Formation 2012
CC-1A
12.24
9.33
0.76
Formation 2012
CC-1A
12.57
9.95
0.79
Formation 2012
CC-1A
12.24
16.85
1.38
Formation 2012
CC-1A
13.55
14.03
1.04
Formation 2012
CC-3A
5.45
10.44
1.92
Formation 2012
CC-3A
5.45
9.20
1.69
Formation 2012
CC-3A
5.48
11.25
2.05
Formation 2012
CC-3A
14.50
15.38
1.06
Formation 2012
CC-3A
14.50
19.68
1.36
Formation 2012
CC-150
4.46
5.83
1.31
Formation 2012
CC-150
4.46
8.67
1.94
Formation 2012
CC-150
4.70
5.20
1.11
Formation 2012
CC-150
7.03
10.14
1.44
Formation 2012
CC-150
14.32
7.83
0.55
Formation 2012
CC-350
3.16
6.28
1.99
Formation 2012
CC-350
3.16
8.53
2.70
Formation 2012
CC-350
4.20
5.78
1.38
Formation 2012
CC-350
11.45
11.50
1.00
Formation 2012
CC-350
11.45
7.95
0.69
Formation 2012
CC-75
3.11
4.05
1.30
Formation 2012
CC-75
3.11
5.35
1.72
Formation 2012
CC-75
3.97
3.18
0.80
Formation 2012
CC-75
4.16
10.32
2.48
Formation 2012
CC-75
4.16
6.60
1.59
Formation 2012
DC-600
8.53
8.54
1.00
Formation 2012
DC-600
8.53
6.20
0.73
Formation 2012
DC-600
8.65
5.85
0.68
Formation 2012
DC-600
7.83
12.83
1.64
B-92

-------
Brown trout (Sulmo trutta)
Study
Site
c
^invert
Cfish
Ratio
Formation 2012
DC-600
7.83
10.54
1.35
Formation 2012
HS
15.70
16.52
1.05
Formation 2012
HS
15.70
25.00
1.59
Formation 2012
HS
18.70
24.90
1.33
Formation 2012
HS
27.80
32.63
1.17
Formation 2012
HS
27.80
22.80
0.82
Formation 2012
HS-3
11.40
20.60
1.81
Formation 2012
HS-3
11.40
18.83
1.65
Formation 2012
HS-3
13.41
17.89
1.33
Formation 2012
HS-3
24.70
23.68
0.96
Formation 2012
HS-3
26.55
28.97
1.09
Formation 2012
LSV-2C
22.62
19.45
0.86
Formation 2012
LSV-2C
22.62
12.78
0.56
Formation 2012
LSV-2C
26.31
22.67
0.86
Formation 2012
LSV-2C
30.00
19.53
0.65
Formation 2012
LSV-2C
26.95
20.96
0.78
Formation 2012
LSV-4
9.54
16.20
1.70
Formation 2012
LSV-4
9.54
15.18
1.59
Formation 2012
SFTC-1
2.42
3.68
1.52
Formation 2012
SFTC-1
3.21
2.25
0.70
Formation 2012
SFTC-1
1.63
6.70
4.11
Formation 2012
SFTC-1
2.49
2.64
1.06
Hamilton and Buhl 2005
CC
6.70
9.70
1.45
McDonald and Strosher 1998
60
ER 747
4.29
4.80
1.12
50
AQ -
0
o
Median ratio:
1.38

o
* 0 e H"
o
O-"""
, ^o ^ O
R2:
F:
df:
0.64
151.8
85

o
0
P:
<0.001

0 S 10 15
20 25 JO 35



.....
B-93

-------
Bullhead (Ameiurus sp.)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1995
ME3
2.55
3.00
1.18
Butler etal. 1993
R2
3.70
3.50
0.95
Butler etal. 1993
R2
3.70
4.00
1.08
Butler etal. 1994
BSW1
5.00
4.10
0.82
fish
2
c
Median ratio: 1.01
R
F
df
P
0.77
6.58
2
0.13
Not used because P > 0.05
invert.
Channel catfish (Ictaluruspunctatus)
Study
Site
Cjnvert
Cfish
Ratio
Butler etal. 1991
7
29.80
21.36
0.72
Butler etal. 1991
7
29.80
22.05
0.74
Butler etal. 1991
7
29.80
17.27
0.58
Butler etal. 1991
7
29.80
19.62
0.66
Butler etal. 1991
7
29.80
22.76
0.76
Butler etal. 1991
7
29.80
24.33
0.82
Butler etal. 1991
7
29.80
32.40
1.09
Butler etal. 1993
LP4
3.20
1.65
0.52
Butler etal. 1993
LP4
3.20
3.30
1.03
Butler etal. 1993
R2
3.90
9.30
2.39
Butler etal. 1993
R2
3.90
1.33
0.34
Butler etal. 1993
R2
3.70
2.04
0.55
Butler etal. 1993
R2
3.70
3.00
0.81
Butler etal. 1995
SJ1
2.50
1.73
0.69
Butler etal. 1995
SJ1
2.50
4.10
1.64
Butler etal. 1995
XT
1.07
1.00
0.94
Butler etal. 1997
MN4
2.65
4.20
1.58
Butler etal. 1997
MN5
8.60
5.00
0.58
Mueller etal. 1991
R1
8.70
2.20
0.25
Roddy et al. 1991
18
3.10
1.40
0.45
Roddy et al. 1991
18
3.10
1.60
0.52
Roddy et al. 1991
18
3.10
1.70
0.55
B-94

-------
Channel catfish (Ictaluruspunctatus)
Study	Site	Cinvert	Cflsh	Ratio
Roddy et al.
1991 18

3.10
1.80
0.58
Roddy et al.
1991 18

3.10
1.90
0.61
Roddy et al.
1991 18

3.10
2.20
0.71
Roddy et al.
1991 18

3.10
1.50
0.48
Roddy et al.
1991 18

3.10
1.70
0.55
Roddy et al.
1991 18

3.10
1.80
0.58
Roddy et al.
1991 18

3.10
2.00
0.65
Roddy et al.
1991 18

3.10
2.10
0.68
Roddy et al.
1991 18

3.10
2.20
0.71
Roddy et al.
1991 18

3.10
2.30
0.74
Roddy et al.
1991 18

3.10
2.40
0.77
Roddy et al.
1991 18
u

3.10
3.10
1.00


0
A
Median ratio: 0.74


20 -
15 
0
0
R2: 0.91





F: 332.8


10 ' O

df: 32


5  0
m a


P: <0.001

II 5 10 15 20 25 3 35
Common carp (Cyprinus carpio)
Study

Site
c
^invert
Cflsh
Ratio
Butler et al.
1991
10
4.80
10.30
2.15
Butler et al.
1991
7
29.80
25.80
0.87
Butler et al.
1991
7
29.80
31.00
1.04
Butler et al.
1991
7
29.80
40.00
1.34
Butler et al.
1991
7
29.80
50.00
1.68
Butler et al.
1991
9
4.10
3.90
0.95
Butler et al.
1991
3
6.20
2.20
0.35
Butler et al.
1993
D2
1.45
3.70
2.55
Butler et al.
1993
F2
7.50
5.80
0.77
Butler et al.
1993
R2
4.30
5.00
1.16
Butler et al.
1993
R2
3.90
4.80
1.23
Butler et al.
1993
R2
3.70
3.30
0.89
Butler et al.
1994
GUN2
28.00
63.00
2.25
Butler et al.
1994
NFK2
3.10
4.90
1.58
B-95

-------
Common carp (Cyprinus carpio)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1994
BSW1
5.00
12.00
2.40
Butler etal. 1994
RBI
21.00
5.10
0.24
Butler etal. 1995
ME4
1.55
3.90
2.52
Butler etal. 1995
ME4
1.55
3.70
2.39
Butler etal. 1995
ME4
1.55
3.80
2.45
Butler etal. 1995
ME3
2.55
4.40
1.73
Butler etal. 1995
ME3
2.55
5.20
2.04
Butler etal. 1995
SJ1
2.50
5.30
2.12
Butler etal. 1995
SJ1
2.50
3.40
1.36
Butler etal. 1995
MN1
2.70
5.80
2.15
Butler etal. 1995
MN1
2.70
9.80
3.63
Butler etal. 1995
MN1
2.70
5.40
2.00
Butler etal. 1997
MN5
8.60
16.00
1.86
Garcia-Hernandez et al. 2000
Cienega de Santa
Clara Wetland
3.00
3.30
1.10
GEI2013
SWB
7.06
12.50
1.77
GEI2013
SWB
7.06
15.61
2.21
GEI 2013
SW11
8.41
3.14
0.37
GEI 2013
SW11
8.41
3.52
0.42
GEI 2013
SW11
8.41
3.66
0.44
GEI 2013
SW11
8.41
3.85
0.46
GEI 2013
SW11
8.41
5.77
0.69
GEI 2013
SW11
8.41
3.60
0.43
GEI 2013
SW11
8.41
3.79
0.45
GEI 2013
SW11
8.41
3.95
0.47
GEI 2013
SW11
8.41
4.14
0.49
GEI 2013
SW11
8.41
4.34
0.52
GEI 2013
SW11
8.41
3.56
0.42
GEI 2013
SW2-1
6.60
26.73
4.05
GEI 2013
SW2-1
6.60
26.74
4.05
GEI 2013
SW2-1
6.60
28.74
4.36
GEI 2013
SW2-1
6.60
29.73
4.51
GEI 2013
SW2-1
6.60
41.57
6.30
GEI 2013
SW2-1
6.60
22.96
3.48
GEI 2013
SW2-1
6.60
24.27
3.68
GEI 2013
SW2-1
6.60
25.09
3.80
GEI 2013
SW2-1
6.60
31.74
4.81
GEI 2013
SW2-1
6.60
36.81
5.58
GEI 2013
SW2-1
9.14
13.29
1.45
GEI 2013
SW2-1
9.14
13.77
1.51
GEI 2013
SW2-1
9.14
20.49
2.24
GEI 2013
SW2-1
9.14
24.84
2.72
B-96

-------
Common carp (Cyprinus carpio)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
SW2-1
9.14
23.65
2.59
GEI2013
SW2-1
9.14
27.27
2.99
GEI 2013
SW4-1
3.33
3.55
1.07
GEI 2013
SW4-1
3.33
4.68
1.41
GEI 2013
SW4-1
3.33
3.91
1.18
GEI 2013
SW4-1
3.33
4.36
1.31
GEI 2013
SW4-1
3.33
4.48
1.35
GEI 2013
SW4-1
3.33
4.60
1.38
GEI 2013
SW4-1
3.33
4.78
1.44
GEI 2013
SW9
4.45
2.73
0.61
GEI 2013
SW9
4.45
2.99
0.67
GEI 2013
SW9
4.45
3.64
0.82
GEI 2013
SW9
4.45
3.80
0.85
GEI 2013
SW9
4.45
3.90
0.88
GEI 2013
SW9
4.45
4.26
0.96
GEI 2013
SW9
4.45
4.53
1.02
GEI 2013
SW9
4.45
3.70
0.83
GEI 2013
SW9
4.45
3.77
0.85
GEI 2013
SW9
4.45
4.14
0.93
GEI 2013
SW9
4.45
4.41
0.99
GEI 2013
SW9
4.45
4.50
1.01
GEI 2013
SW9
4.45
4.69
1.05
GEI 2013
SW88
3.96
3.88
0.98
GEI 2013
SW88
3.96
5.33
1.35
GEI 2013
SW88
3.96
5.49
1.39
GEI 2013
SW88
3.96
5.66
1.43
Grassoetal. 1995
9
7.59
4.70
0.62
Grassoetal. 1995
9
7.59
4.93
0.65
Grassoetal. 1995
9
7.59
5.51
0.73
Lambing et al. 1994
S34
14.00
19.00
1.36
Lambing et al. 1994
S34
14.00
32.00
2.29
Low and Mullins 1990
5
5.60
1.20
0.21
Low and Mullins 1990
7
1.60
0.30
0.19
May et al. 2008
KR
17.20
7.78
0.45
May et al. 2008
NSCL
10.70
10.80
1.01
May et al. 2008
NSK
8.81
9.33
1.06
May et al. 2008
NSP
24.00
10.30
0.43
May et al. 2008
SSAL
11.50
10.50
0.91
May et al. 2008
SSAU
8.35
7.59
0.91
May et al. 2008
sso
10.00
8.48
0.85
May et al. 2008
ssw
7.60
10.40
1.37
Mueller etal. 1991
R2
6.40
14.40
2.25
B-97

-------
Common carp (Cyprinus carpio)
Study
Site
r
^invert
Cflsh
Ratio
Mueller etal. 1991
R2
6.40
14.00
2.19
Mueller etal. 1991
R1
8.70
5.60
0.64
Mueller etal. 1991
A3
6.00
6.50
1.08
Mueller etal. 1991
A6
5.60
3.40
0.61
Mueller etal. 1991
A2
8.50
7.30
0.86
Peterson et al. 1991
7
3.83
4.24
1.11
Peterson et al. 1991
7
3.83
4.41
1.15
Peterson et al. 1991
7
3.83
4.73
1.23
Peterson et al. 1991
7
3.83
5.16
1.35
Peterson et al. 1991
7
3.83
5.21
1.36
Roddy et al. 1991
18
3.10
3.20
1.03
Roddy et al. 1991
18
3.10
3.90
1.26
Roddy et al. 1991
18
3.10
4.60
1.48
Roddy et al. 1991
18
3.10
4.70
1.52
Roddy et al. 1991
18
3.10
4.80
1.55
Roddy et al. 1991
18
3.10
5.30
1.71
Lemly 1985
Badin Lake
5.70
3.17
0.56
Lemly 1985
Belews Lake
51.15
21.29
0.42
Lemly 1985
High Rock Lake
9.05
2.45
0.27
Rinellaand Schuler 1992
Harney Lake
2.05
2.20
1.07
Rinellaand Schuler 1992
S. Malheur Lake
1.20
2.00
1.67
o
	Median ratio: 1.20
0	10 20 30 40 50 60
^inverts
Creek chub (Semotilus atromaculatus)
Study	Site	Cinvert	Cflsh	Ratio
Mason et al. 2000	HCRT	2.81	0.49	0.18
Mason et al. 2000	HCRT	2.81	1.18	0.42
Mason et al. 2000	HCRT	2.81	1.97	0.70
GEI2013	SW4-1	3.33	4.65	1.40
GEI2013	SW4-1	3.33	4.96	1.49
B-98

-------
Creek chub (Semotilus atromaculatus)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
SW4-1
3.33
5.52
1.66
GEI2013
SW4-1
3.33
6.11
1.84
GEI 2013
SW4-1
3.33
6.31
1.90
GEI 2013
SW4-1
3.33
6.53
1.96
GEI 2013
SW4-1
3.33
6.67
2.01
GEI 2013
LG1
3.37
3.41
1.01
GEI 2013
LG1
3.37
3.58
1.06
GEI 2013
LG1
3.37
3.75
1.11
GEI 2013
LG1
3.37
3.78
1.12
GEI 2013
LG1
3.37
4.10
1.22
GEI 2013
LG1
3.39
3.23
0.95
GEI 2013
LG1
3.39
3.72
1.10
GEI 2013
LG1
3.39
3.74
1.10
GEI 2013
LG1
3.39
3.78
1.12
GEI 2013
LG1
3.39
3.89
1.15
GEI 2013
LG1
3.39
4.03
1.19
GEI 2013
LG1
3.39
4.12
1.22
GEI 2013
LG1
3.39
5.11
1.51
GEI 2013
LG1
3.39
5.21
1.54
GEI 2013
LG1
3.39
5.34
1.58
GEI 2013
LG1
3.56
3.28
0.92
GEI 2013
LG1
3.56
3.37
0.95
GEI 2013
LG1
3.56
3.82
1.07
GEI 2013
LG1
3.56
3.86
1.09
GEI 2013
LG1
3.56
4.02
1.13
GEI 2013
LG1
3.56
4.16
1.17
GEI 2013
LG1
3.56
4.49
1.26
GEI 2013
LG1
3.56
4.53
1.27
GEI 2013
LG1
3.56
4.63
1.30
GEI 2013
LG1
3.56
4.77
1.34
GEI 2013
CC1
3.76
5.43
1.44
GEI 2013
CC1
3.76
5.57
1.48
GEI 2013
CC1
3.76
6.51
1.73
GEI 2013
CC1
3.76
6.71
1.78
GEI 2013
CC1
3.76
7.12
1.89
GEI 2013
CC1
4.69
3.99
0.85
GEI 2013
CC1
4.69
4.06
0.87
GEI 2013
CC1
4.69
4.08
0.87
GEI 2013
CC1
4.69
4.25
0.91
GEI 2013
CC1
4.69
4.44
0.95
GEI 2013
CC1
4.69
4.48
0.96
GEI 2013
CC1
4.69
4.50
0.96
B-99

-------
Creek chub (Semotilus atromaculatus)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
CC1
4.69
4.72
1.01
GEI2013
CC1
4.69
5.24
1.12
GEI 2013
CC1
4.69
5.44
1.16
GEI 2013
CC1
5.86
4.98
0.85
GEI 2013
CC1
5.86
5.39
0.92
GEI 2013
CC1
5.86
5.77
0.99
GEI 2013
CC1
5.86
6.39
1.09
GEI 2013
CC1
5.86
6.43
1.10
GEI 2013
CC1
5.86
6.50
1.11
GEI 2013
CC1
5.86
6.57
1.12
GEI 2013
CC1
5.86
7.42
1.27
GEI 2013
CC1
5.86
7.42
1.27
GEI 2013
CC1
5.86
7.47
1.28
GEI 2014
Bond Creek, BC-2
2.96
2.46
0.83
GEI 2014
Bond Creek, BC-2
2.96
3.22
1.09
GEI 2014
Bond Creek, BC-2
2.96
2.64
0.89
GEI 2014
Bond Creek, BC-2
2.96
2.96
1.00
GEI 2014
Bond Creek, BC-2
2.96
4.47
1.51
GEI 2014
Bond Creek, BC-3
3.02
3.09
1.02
GEI 2014
Bond Creek, BC-3
3.02
2.91
0.96
GEI 2014
Bond Creek, BC-3
3.02
3.45
1.14
GEI 2014
Bond Creek, BC-3
3.02
2.69
0.89
GEI 2014
Bond Creek, BC-3
3.02
3.30
1.09
GEI 2014
Bond Creek, BC-3
5.87
3.44
0.59
GEI 2014
Bond Creek, BC-3
5.87
2.62
0.45
GEI 2014
Bond Creek, BC-3
5.87
3.23
0.55
GEI 2014
Cow Camp Creek, CC-2
5.65
3.05
0.54
GEI 2014
Cow Camp Creek, CC-2
5.65
3.69
0.65
GEI 2014
Cow Camp Creek, CC-2
5.65
3.84
0.68
GEI 2014
Cow Camp Creek, CC-2
5.65
4.44
0.79
GEI 2014
Cow Camp Creek, CC-2
5.65
3.98
0.70
GEI 2014
Hazy Creek, C-HC1
8.03
6.84
0.85
GEI 2014
Hazy Creek, C-HC1
8.03
3.78
0.47
GEI 2014
Hazy Creek, C-HC1
8.03
5.81
0.72
GEI 2014
Hazy Creek, C-HC1
8.03
4.29
0.53
GEI 2014
Hazy Creek, C-HC1
8.03
3.59
0.45
GEI 2014
Laurel Fork, C-LF1
12.73
6.52
0.51
GEI 2014
Laurel Fork, C-LF1
12.73
6.81
0.54
GEI 2014
Laurel Fork, C-LF1
12.73
5.11
0.40
GEI 2014
Laurel Fork, C-LF1
12.73
5.16
0.41
GEI 2014
Laurel Fork, C-LF1
12.73
5.46
0.43
GEI 2014
Little Marsh Fork, C-LMF1
6.02
3.81
0.63
B-100

-------
Creek chub (Semotilus atromaculatus)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
Little Marsh Fork, C-LMF1
6.02
4.89
0.81
GEI2014
Little Marsh Fork, C-LMF 1
6.02
3.58
0.60
GEI 2014
Little Marsh Fork, C-LMF 1
6.02
4.81
0.80
GEI 2014
Little Marsh Fork, C-LMF 1
6.02
5.82
0.97
GEI 2014
Dry Creek, DC-1
3.37
4.62
1.37
GEI 2014
Dry Creek, DC-1
3.37
5.15
1.53
GEI 2014
Dry Creek, DC-1
3.37
6.46
1.92
GEI 2014
Dry Creek, DC-1
3.37
5.73
1.70
GEI 2014
Dry Creek, DC-1
3.37
4.59
1.36
GEI 2014
Dry Creek, DC-1
2.29
4.94
2.16
GEI 2014
Dry Creek, DC-1
2.29
4.04
1.76
GEI 2014
Dry Creek, DC-1
2.29
3.62
1.58
GEI 2014
Dry Creek, DC-1
2.29
3.84
1.68
GEI 2014
Dry Creek, DC-1
2.15
4.02
1.87
GEI 2014
Dry Creek, DC-1
2.15
3.60
1.67
GEI 2014
Dry Creek, DC-1
2.15
3.28
1.52
GEI 2014
Dry Creek, DC-1
2.15
3.03
1.41
GEI 2014
Dry Creek, DC-1
2.15
4.01
1.86
GEI 2014
Dry Creek, DC-1
1.97
3.51
1.78
GEI 2014
Dry Creek, DC-2
3.42
5.32
1.56
GEI 2014
Dry Creek, DC-2
3.42
4.62
1.35
GEI 2014
Dry Creek, DC-2
3.42
4.43
1.30
GEI 2014
Dry Creek, DC-2
3.42
4.56
1.34
GEI 2014
Dry Creek, DC-2
3.42
6.38
1.87
GEI 2014
Dry Creek, DC-2
3.42
2.96
0.87
GEI 2014
Dry Creek, DC-2
3.42
3.58
1.05
GEI 2014
Dry Creek, DC-2
3.42
3.22
0.94
GEI 2014
Dry Creek, DC-2
3.42
4.07
1.19
GEI 2014
Dry Creek, DC-2
3.42
3.28
0.96
GEI 2014
Dry Creek, DC-2
3.16
6.52
2.07
GEI 2014
Dry Creek, DC-2
3.16
4.92
1.56
GEI 2014
Dry Creek, DC-2
3.16
3.10
0.98
GEI 2014
Dry Creek, DC-2
3.16
3.14
1.00
GEI 2014
Dry Creek, DC-2
3.16
4.36
1.38
GEI 2014
Dry Creek, DC-2
3.16
3.12
0.99
GEI 2014
Dry Creek, DC-2
3.16
5.40
1.71
GEI 2014
Dry Creek, DC-2
2.93
2.85
0.97
GEI 2014
Dry Creek, DC-2
2.93
4.94
1.69
GEI 2014
Dry Creek, DC-2
2.93
5.17
1.76
GEI 2014
Dry Creek, DC-2
2.93
3.47
1.18
GEI 2014
Dry Creek, DC-2
2.93
2.49
0.85
GEI 2014
Dry Creek, DC-3
7.18
23.79
3.31
B-101

-------
Creek chub (Semotilus atromaculatus)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
Dry Creek, DC-3
7.18
16.06
2.24
GEI2014
Dry Creek, DC-3
7.18
24.43
3.40
GEI 2014
Dry Creek, DC-3
7.18
22.20
3.09
GEI 2014
Dry Creek, DC-3
7.18
26.28
3.66
GEI 2014
Dry Creek, DC-3
9.23
21.48
2.33
GEI 2014
Dry Creek, DC-3
9.23
21.24
2.30
GEI 2014
Dry Creek, DC-3
9.23
21.46
2.33
GEI 2014
Dry Creek, DC-3
9.23
22.48
2.44
GEI 2014
Dry Creek, DC-3
9.23
18.80
2.04
GEI 2014
Dry Creek, DC-3
9.23
16.87
1.83
GEI 2014
Dry Creek, DC-4
19.42
15.86
0.82
GEI 2014
Dry Creek, DC-4
19.42
12.76
0.66
GEI 2014
Dry Creek, DC-4
19.42
28.50
1.47
GEI 2014
Dry Creek, DC-4
19.42
18.14
0.93
GEI 2014
Dry Creek, DC-4
19.42
17.55
0.90
GEI 2014
Dry Creek, DC-4
18.10
34.60
1.91
GEI 2014
Dry Creek, DC-4
18.10
23.70
1.31
GEI 2014
Foidel Creek, FOC-1
3.06
8.00
2.61
GEI 2014
Foidel Creek, FOC-1
3.06
9.68
3.16
GEI 2014
Foidel Creek, FOC-1
3.06
8.86
2.89
GEI 2014
Foidel Creek, FOC-1
3.06
2.51
0.82
GEI 2014
Foidel Creek, FOC-1
3.06
2.86
0.93
GEI 2014
Foidel Creek, FOC-1
3.06
4.24
1.39
GEI 2014
Foidel Creek, FOC-1
3.06
3.27
1.07
GEI 2014
Foidel Creek, FOC-1
3.06
5.03
1.64
GEI 2014
Foidel Creek, FOC-2
2.18
2.07
0.95
GEI 2014
Foidel Creek, FOC-2
2.18
3.06
1.41
GEI 2014
Foidel Creek, FOC-2
2.18
3.82
1.76
GEI 2014
Foidel Creek, FOC-2
2.18
2.26
1.04
GEI 2014
Foidel Creek, FOC-2
2.18
2.02
0.93
GEI 2014
Foidel Creek, FOC-2
2.18
2.28
1.05
GEI 2014
Foidel Creek, FOC-2
2.18
2.44
1.12
GEI 2014
Foidel Creek, FOC-2
2.18
2.62
1.21
GEI 2014
Grassy Creek, GC-2
4.20
5.28
1.26
GEI 2014
Grassy Creek, GC-2
4.20
6.13
1.46
GEI 2014
Grassy Creek, GC-2
4.20
6.29
1.50
GEI 2014
Grassy Creek, GC-2
4.20
4.80
1.15
GEI 2014
Grassy Creek, GC-2
4.20
4.59
1.09
GEI 2014
Grassy Creek, GC-2
4.58
3.27
0.71
GEI 2014
Grassy Creek, GC-2
4.58
5.50
1.20
GEI 2014
Grassy Creek, GC-2
4.58
3.64
0.80
GEI 2014
Grassy Creek, GC-2
4.58
4.29
0.94
B-102

-------
Creek chub (Semotilus atromaculatus)
Study
Site

r
^invert
Cfish
Ratio
GEI2014
Grassy Creek
GC-2
4.58
3.01
0.66
GEI2014
Grassy Creek
GC-2
3.96
7.48
1.89
GEI 2014
Grassy Creek
GC-2
3.96
6.12
1.55
GEI 2014
Grassy Creek
GC-2
3.96
8.61
2.18
GEI 2014
Grassy Creek
GC-2
3.96
7.09
1.79
GEI 2014
Grassy Creek
GC-2
3.96
5.06
1.28
GEI 2014
Grassy Creek
GC-2
3.96
4.95
1.25
GEI 2014
Grassy Creek
GC-2
3.96
3.85
0.97
GEI 2014
Grassy Creek
GC-2
3.96
5.32
1.34
GEI 2014
Grassy Creek
GC-2
3.96
4.04
1.02
GEI 2014
Grassy Creek
GC-2
3.96
4.25
1.08
GEI 2014
Grassy Creek
GC-2
3.97
3.48
0.88
GEI 2014
Grassy Creek
GC-2
3.97
3.86
0.97
GEI 2014
Grassy Creek
GC-2
3.97
4.08
1.03
GEI 2014
Grassy Creek
GC-2
3.97
4.58
1.15
GEI 2014
Grassy Creek
GC-2
3.97
3.53
0.89
GEI 2014
Grassy Creek
GC-3
4.54
6.07
1.34
GEI 2014
Grassy Creek
GC-3
4.54
7.25
1.60
GEI 2014
Grassy Creek
GC-3
4.54
5.25
1.16
GEI 2014
Grassy Creek
GC-3
4.54
5.65
1.25
GEI 2014
Grassy Creek
GC-3
4.54
10.75
2.37
GEI 2014
Grassy Creek
GC-3
4.55
3.30
0.73
GEI 2014
Grassy Creek
GC-3
4.55
3.98
0.88
GEI 2014
Grassy Creek
GC-3
4.55
3.46
0.76
GEI 2014
Grassy Creek
GC-3
4.55
4.14
0.91
GEI 2014
Grassy Creek
GC-3
4.55
3.81
0.84
GEI 2014
Grassy Creek
GC-3
4.34
11.05
2.55
GEI 2014
Grassy Creek
GC-3
4.34
7.22
1.66
GEI 2014
Grassy Creek
GC-3
4.34
9.61
2.22
GEI 2014
Grassy Creek
GC-3
4.34
6.03
1.39
GEI 2014
Grassy Creek
GC-3
4.34
3.95
0.91
GEI 2014
Grassy Creek
GC-3
4.34
4.82
1.11
GEI 2014
Grassy Creek
GC-3
4.34
5.18
1.19
GEI 2014
Grassy Creek
GC-3
4.34
4.46
1.03
GEI 2014
Grassy Creek
GC-3
4.34
4.21
0.97
GEI 2014
Grassy Creek
GC-3
4.35
2.97
0.68
GEI 2014
Grassy Creek
GC-3
4.35
4.16
0.96
GEI 2014
Grassy Creek
GC-3
4.35
4.29
0.99
GEI 2014
Grassy Creek
GC-3
4.35
3.69
0.85
GEI 2014
Grassy Creek
GC-3
4.35
4.73
1.09
GEI 2014
Grassy Creek
GC-4
5.10
4.30
0.84
GEI 2014
Grassy Creek
GC-4
5.10
5.17
1.01
B-103

-------
Creek chub (Semotilus atromaculatus)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
Grassy Creek, GC-4
5.10
5.46
1.07
GEI2014
Grassy Creek, GC-4
5.10
5.63
1.10
GEI 2014
Grassy Creek, GC-4
5.10
5.15
1.01
GEI 2014
Grassy Creek, GC-4
5.76
7.72
1.34
GEI 2014
Grassy Creek, GC-4
5.76
6.04
1.05
GEI 2014
Grassy Creek, GC-4
5.76
7.88
1.37
GEI 2014
Grassy Creek, GC-4
5.76
9.77
1.70
GEI 2014
Grassy Creek, GC-4
5.76
7.35
1.28
GEI 2014
Grassy Creek, GC-4
5.76
4.17
0.72
GEI 2014
Grassy Creek, GC-4
5.76
4.86
0.84
GEI 2014
Grassy Creek, GC-4
5.76
5.02
0.87
GEI 2014
Grassy Creek, GC-4
5.76
4.79
0.83
GEI 2014
Grassy Creek, GC-4
5.76
6.56
1.14
GEI 2014
Grassy Creek, GC-4
5.76
5.31
0.92
GEI 2014
Grassy Creek, GC-4 US
13.16
4.05
0.31
GEI 2014
Grassy Creek, GC-4 US
13.16
4.80
0.36
GEI 2014
Grassy Creek, GC-4 US
13.16
5.41
0.41
GEI 2014
Grassy Creek, GC-4 US
13.16
6.05
0.46
GEI 2014
Big Horse Creek, H-BHC3
5.78
3.96
0.69
GEI 2014
Big Horse Creek, H-BHC3
5.78
2.97
0.51
GEI 2014
Big Horse Creek, H-BHC3
5.78
3.84
0.66
GEI 2014
Sally Fork, H-BLB2
1.64
2.42
1.48
GEI 2014
Sally Fork, H-BLB2
1.64
1.65
1.01
GEI 2014
Sally Fork, H-BLB2
1.64
1.68
1.03
GEI 2014
Sally Fork, H-BLB2
1.64
2.02
1.23
GEI 2014
Sally Fork, H-BLB2
1.64
1.46
0.89
GEI 2014
Hubberson Gulch, HG-2
3.44
4.41
1.28
GEI 2014
Hubberson Gulch, HG-2
3.44
3.56
1.04
GEI 2014
Hubberson Gulch, HG-2
3.44
4.48
1.30
GEI 2014
Hubberson Gulch, HG-2
1.46
2.95
2.03
GEI 2014
Hubberson Gulch, HG-2
1.46
2.66
1.83
GEI 2014
Hubberson Gulch, HG-2
Jack Smith (Bear) Branch,
1.46
2.87
1.97
GEI 2014
H-JSB1
Jack Smith (Bear) Branch,
4.03
3.04
0.75
GEI 2014
H-JSB1
Jack Smith (Bear) Branch,
4.03
1.81
0.45
GEI 2014
H-JSB1
Jack Smith (Bear) Branch,
4.03
2.35
0.58
GEI 2014
H-JSB1
Jack Smith (Bear) Branch,
4.03
1.91
0.47
GEI 2014
H-JSB1
4.03
2.83
0.70
GEI 2014
Laurel Creek, H-LC1
4.57
1.29
0.28
B-104

-------
Creek chub (Semotilus atromaculatus)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
Laurel Creek, H-LC1
4.57
2.04
0.45
GEI2014
Laurel Creek, H-LC1
4.57
1.49
0.33
GEI 2014
Laurel Creek, H-LC1
4.57
1.85
0.41
GEI 2014
Laurel Creek, H-LC1
4.57
0.67
0.15
GEI 2014
Lick Creek, H-LKC1
2.59
1.83
0.71
GEI 2014
Lick Creek, H-LKC 1
2.59
1.40
0.54
GEI 2014
Lick Creek, H-LKC 1
2.59
1.41
0.54
GEI 2014
Lick Creek, H-LKC 1
2.59
1.19
0.46
GEI 2014
Lick Creek, H-LKC 1
2.59
1.22
0.47
GEI 2014
Mud River, H-MR3
3.86
4.75
1.23
GEI 2014
Mud River, H-MR3
3.86
4.60
1.19
GEI 2014
Mud River, H-MR3
3.86
5.06
1.31
GEI 2014
Mud River, H-MR3
3.86
3.32
0.86
GEI 2014
Mud River, H-MR3
3.86
4.19
1.08
GEI 2014
Mud River, H-MR5
3.58
1.51
0.42
GEI 2014
Mud River, H-MR5
3.58
1.43
0.40
GEI 2014
Mud River, H-MR5
3.58
1.98
0.55
GEI 2014
Mud River, H-MR5
3.58
3.80
1.06
GEI 2014
Mud River, H-MR5
3.58
3.44
0.96
GEI 2014
Sugartree Branch, H-SB1
10.62
7.29
0.69
GEI 2014
Sugartree Branch, H-SB 1
10.62
7.56
0.71
GEI 2014
Sugartree Branch, H-SB 1
10.62
6.20
0.58
GEI 2014
Middle Creek, MC-1
3.21
2.75
0.86
GEI 2014
Middle Creek, MC-1
3.21
4.74
1.48
GEI 2014
Middle Creek, MC-1
3.21
4.01
1.25
GEI 2014
Middle Creek, MC-1
3.21
3.94
1.23
GEI 2014
Middle Creek, MC-1
3.21
3.63
1.13
GEI 2014
Middle Creek, MC-1
3.21
1.83
0.57
GEI 2014
Middle Creek, MC-1
3.21
1.85
0.57
GEI 2014
Middle Creek, MC-1
3.21
2.50
0.78
GEI 2014
Middle Creek, MC-1
3.21
2.33
0.73
GEI 2014
Middle Creek, MC-1
3.21
2.56
0.80
GEI 2014
Middle Creek, MC-2
4.19
1.93
0.46
GEI 2014
Middle Creek, MC-2
4.19
1.89
0.45
GEI 2014
Middle Creek, MC-2
4.19
2.51
0.60
GEI 2014
Middle Creek, MC-2
4.19
1.87
0.45
GEI 2014
Middle Creek, MC-2
4.19
2.13
0.51
GEI 2014
Sage Creek, SC-3
2.29
7.33
3.20
GEI 2014
Sage Creek, SC-3
2.29
7.23
3.16
GEI 2014
Sage Creek, SC-3
2.29
7.60
3.32
GEI 2014
Sage Creek, SC-3
2.29
4.77
2.08
GEI 2014
Sage Creek, SC-3
2.29
8.70
3.80
B-105

-------
Creek chub (Semotilus atromaculatus)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
Sage Creek, SC-4
23.79
18.77
0.79
GEI2014
Sage Creek, SC-4
23.79
20.08
0.84
GEI 2014
Sage Creek, SC-4
23.79
13.05
0.55
GEI 2014
Scotchmans Gulch, SG-1A
7.16
7.41
1.04
GEI 2014
Scotchmans Gulch, SG-1A
7.16
7.65
1.07
GEI 2014
Scotchmans Gulch, SG-1A
7.16
4.47
0.62
GEI 2014
Scotchmans Gulch, SG-1A
7.16
5.83
0.81
GEI 2014
Scotchmans Gulch, SG-1A
7.16
5.37
0.75
GEI 2014
Scotchmans Gulch, SG-1A
7.16
5.69
0.80
GEI 2014
Scotchmans Gulch, SG-1A
7.16
4.31
0.60
o
Median ratio: 1.06
10	15
^inverts
o
20
25
R2:
0.41
F:
214.7
df:
303
P:
<0.001
Cutthroat trout (Oncorhynchus clarkii)
Study
Site
r
^invert
Cfish
Ratio
Hamilton and Buhl 2004
ShpC
1.90
1.80
0.95
McDonald and Strosher 1998
ER 745
2.74
5.40
1.97
Hamilton and Buhl 2005
SC
4.10
3.50
0.85
McDonald and Strosher 1998
ER 747
4.29
6.57
1.53
Hamilton and Buhl 2005
UAC
5.00
6.60
1.32
Hamilton and Buhl 2004
ACM
6.70
6.30
0.94
Hamilton and Buhl 2005
DC
8.70
11.00
1.26
McDonald and Strosher 1998
ER 746
10.70
12.71
1.19
Hamilton and Buhl 2005
BGS
10.80
12.20
1.13
Hamilton and Buhl 2004
DVC
12.80
10.20
0.80
Hamilton and Buhl 2004
UEMC
26.90
27.00
1.00
Hamilton and Buhl 2004
LEMC
75.20
52.30
0.70
Minnow 2007
BA6
3.27
6.98
2.13
Minnow 2007
AL4
3.92
4.44
1.13
Minnow 2007
MI5
4.00
5.12
1.28
B-106

-------
Cutthroat trout (Oncorhynchus clarkii)
Study
Site
c
^invert
Cfish
Ratio
Minnow 2007
EL12
4.01
7.42
1.85
Minnow 2007
EL14
4.41
4.52
1.02
Minnow 2007
F09
4.44
7.80
1.76
Minnow 2007
MI3
6.21
5.65
0.91
Minnow 2007
MI2
6.69
5.16
0.77
Minnow 2007
ELI
7.08
4.82
0.68
Minnow 2007
LI8
7.81
9.36
1.20
Minnow 2007
FO10
17.51
45.94
2.62
Minnow 2007
HA7
22.41
21.10
0.94
Minnow 2007
CL11
30.87
57.27
1.86
Orr et al. 2012
Alexander Creek
3.92
2.71
0.69
Orr et al. 2012
Elk River 1
6.23
5.61
0.90
Orr et al. 2012
Elk River 1
6.23
7.94
1.27
Orr et al. 2012
Elk River 1
7.08
5.35
0.76
Orr et al. 2012
Elk River 12
3.81
4.58
1.20
Orr et al. 2012
Elk River 12
4.01
4.89
1.22
Orr et al. 2012
Fording River 22
3.10
10.28
3.32
Orr et al. 2012
Fording River 23
7.72
7.92
1.03
Orr et al. 2012
Fording River 9
4.44
6.92
1.56
Orr et al. 2012
Line Creek 8
6.61
7.99
1.21
Orr et al. 2012
Line Creek 8
7.81
7.13
0.91
Orr et al. 2012
Michel Creek 2
8.38
5.13
0.61
Orr et al. 2012
Michel Creek 2
6.69
4.63
0.69
Orr et al. 2012
Michel Creek 3
5.42
3.51
0.65
Orr et al. 2012
Michel Creek 3
6.21
4.02
0.65
Orr et al. 2012
Michel Creek 5
4.00
4.12
1.03
Orr et al. 2012
Fording River
MP1
Barnes Lake
5.49
6.84
1.25
Orr et al. 2012
Wetland 6
3.27
3.92
1.20
Orr et al. 2012
Clode Pond 11
32.22
41.27
1.28
Orr et al. 2012
Clode Pond 11
30.87
44.70
1.45
Orr et al. 2012
Elk Lakes 14
6.40
7.14
1.12
Orr et al. 2012
Elk Lakes 14
4.41
4.35
0.99
Orr et al. 2012
Fording River
Oxbow 10
49.26
24.34
0.49
Orr et al. 2012
Fording River
Oxbow 10
17.51
34.41
1.97
Orr et al. 2012
Henretta Lake 27
9.16
6.90
0.75
Orr et al. 2012
O'Rourke Lake 1
3.63
8.05
2.22
Orr et al. 2012
Harmer Pond 7
22.41
13.08
0.58
B-107

-------
Cutthroat trout (Oncorhynchus clarkii)
Study Site
c
^invert
Cfish
Ratio
60
50
40
30
20
10
0
O
O
O
%
a-
^0'
o
20
40
^inverts
60
Median ratio: 1.12
R
F
df
P
0.66
95.81
50
<0.001
Fathead minnow (Pimephales promelas)
Study
Site
Cjnvert
Cfish
Ratio
Birkner 1978
4
1.80
2.10
1.17
Birkner 1978
22
11.30
11.00
0.97
Birkner 1978
27
34.60
79.00
2.28
Birkner 1978
23
15.50
34.50
2.23
Birkner 1978
1
1.75
2.10
1.20
Butler etal. 1991
10
4.80
8.10
1.69
Butler etal. 1991
3
6.20
9.50
1.53
Butler etal. 1993
SP2
3.40
6.00
1.76
Butler etal. 1993
SP2
3.15
8.20
2.60
Butler etal. 1993
D1
1.20
3.70
3.08
Butler etal. 1993
D1
1.20
3.80
3.17
Butler etal. 1993
U1
2.45
6.40
2.61
Butler etal. 1993
R2
3.90
6.60
1.69
Butler etal. 1993
R2
3.70
6.60
1.78
Butler etal. 1993
ST2
4.50
12.80
2.84
Butler etal. 1993
ST2
4.10
7.60
1.85
Butler etal. 1993
ST2
4.10
16.00
3.90
Butler etal. 1993
R1
4.00
11.00
2.75
Butler etal. 1993
R1
4.00
11.00
2.75
Butler etal. 1993
SB2
3.75
5.70
1.52
Butler etal. 1993
SB2
3.75
8.60
2.29
Butler etal. 1993
SB2
3.65
9.90
2.71
Butler etal. 1993
WSB2
4.75
17.10
3.60
Butler etal. 1993
WSB2
3.60
4.20
1.17
Butler etal. 1993
WSB2
3.60
10.00
2.78
B-108

-------
Fathead minnow (Pimephales promelas)
Study

Site
c
^invert
Cfish
Ratio
Butler et al.
1993
WSB2
3.00
8.10
2.70
Butler et al.
1994
CRC
7.50
20.40
2.72
Butler et al.
1994
CF1
3.60
7.90
2.19
Butler et al.
1994
GUN2
28.00
7.50
0.27
Butler et al.
1994
IW
8.35
10.00
1.20
Butler et al.
1994
TGC
4.90
11.00
2.24
Butler et al.
1994
AD
2.70
9.60
3.56
Butler et al.
1994
LSW1
3.90
73.00
18.72
Butler et al.
1994
OMD
73.00
13.00
0.18
Butler et al.
1994
PSW1
3.70
22.00
5.95
Butler et al.
1994
MKP
32.00
51.00
1.59
Butler et al.
1995
AK
0.78
2.60
3.35
Butler et al.
1995
AK
0.78
2.90
3.74
Butler et al.
1995
AK
0.78
2.80
3.61
Butler et al.
1995
DD
0.86
3.40
3.95
Butler et al.
1995
DD
0.86
3.90
4.53
Butler et al.
1995
DD
0.86
3.60
4.19
Butler et al.
1995
HD1
0.83
3.90
4.73
Butler et al.
1995
HD1
0.83
2.50
3.03
Butler et al.
1995
HD1
0.83
2.60
3.15
Butler et al.
1995
HD2
0.98
1.50
1.53
Butler et al.
1995
HD2
0.98
1.60
1.63
Butler et al.
1995
ME1
3.40
5.60
1.65
Butler et al.
1995
ME2
1.25
4.80
3.84
Butler et al.
1995
ME4
1.55
1.40
0.90
Butler et al.
1995
ME4
1.55
5.90
3.81
Butler et al.
1995
ME3
2.55
4.30
1.69
Butler et al.
1995
ME3
2.55
5.30
2.08
Butler et al.
1995
ME3
2.55
4.40
1.73
Butler et al.
1995
SD
1.40
4.90
3.50
Butler et al.
1995
SD
1.40
3.00
2.14
Butler et al.
1995
SD
1.40
4.00
2.86
Butler et al.
1995
WC
6.75
18.40
2.73
Butler et al.
1995
WC
6.75
22.90
3.39
Butler et al.
1995
WC
6.75
26.40
3.91
Butler et al.
1995
YJ2
1.65
11.00
6.67
Butler et al.
1995
YJ2
1.65
4.00
2.42
Butler et al.
1997
MNP2
4.40
11.00
2.50
Butler et al.
1997
MUD2
3.45
7.70
2.23
Butler et al.
1997
MUD2
3.45
12.00
3.48
Butler et al.
1997
MUD2
3.45
6.50
1.88
Butler et al.
1997
WCP
9.70
10.00
1.03
B-109

-------
Fathead minnow (Pimephales promelas)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1997
WCP
9.70
15.00
1.55
Butler etal. 1997
TR25
1.80
4.00
2.22
Butler etal. 1997
TR25
1.80
5.20
2.89
Butler etal. 1997
TR25
1.80
6.00
3.33
Butler etal. 1997
TRH
1.60
4.20
2.63
Butler etal. 1997
TRH
1.60
4.30
2.69
Butler etal. 1997
TRH
1.60
2.20
1.38
Butler etal. 1997
TRH
1.60
3.00
1.88
Butler etal. 1997
MN5
8.60
7.30
0.85
Butler etal. 1997
MNP1
0.70
1.70
2.43
Butler etal. 1997
MNP1
0.70
1.80
2.57
GEI2013
SWA1
3.64
4.07
1.12
GEI2013
SWA1
3.64
4.68
1.29
GEI 2013
SWA1
3.64
4.76
1.31
GEI 2013
SWA1
3.64
5.45
1.50
GEI 2013
SWA1
3.64
5.71
1.57
GEI 2013
SWA1
3.64
3.62
1.00
GEI 2013
SWA1
3.64
3.72
1.02
GEI 2013
SWA1
3.64
4.43
1.22
GEI 2013
SWA1
3.64
4.52
1.24
GEI 2013
SWA1
3.64
4.66
1.28
GEI 2013
SWA1
2.81
4.48
1.60
GEI 2013
SWA1
2.81
4.53
1.61
GEI 2013
SWA1
2.81
5.00
1.78
GEI 2013
SWA1
2.81
5.24
1.87
GEI 2013
SWA1
2.81
5.76
2.05
GEI 2013
SWA1
2.81
3.89
1.39
GEI 2013
SWA1
2.81
3.98
1.42
GEI 2013
SWA1
2.81
4.04
1.44
GEI 2013
SWA1
2.81
4.33
1.54
GEI 2013
SWA1
2.81
4.81
1.71
GEI 2013
SWB
7.06
7.38
1.05
GEI 2013
SWB
7.06
8.49
1.20
GEI 2013
SWB
7.06
8.72
1.24
GEI 2013
SWB
7.06
9.80
1.39
GEI 2013
SWB
7.06
8.61
1.22
GEI 2013
SWB
7.06
9.02
1.28
GEI 2013
SWB
7.06
9.11
1.29
GEI 2013
SWB
7.06
9.30
1.32
GEI 2013
SWB
7.06
9.53
1.35
GEI 2013
SWB
7.44
10.97
1.48
GEI 2013
SWB
7.44
11.22
1.51
B-110

-------
Fathead minnow (Pimephales promelas)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
SWB
7.44
12.25
1.65
GEI2013
SWB
7.44
12.43
1.67
GEI 2013
SWB
7.44
12.46
1.68
GEI 2013
SWB
7.44
9.36
1.26
GEI 2013
SWB
7.44
9.46
1.27
GEI 2013
SWB
7.44
9.78
1.32
GEI 2013
SWB
7.44
9.87
1.33
GEI 2013
SWB
7.44
10.66
1.43
GEI 2013
SW11
8.41
5.70
0.68
GEI 2013
SW11
8.41
7.05
0.84
GEI 2013
SW11
8.41
5.38
0.64
GEI 2013
SW11
8.41
4.68
0.56
GEI 2013
SW11
8.41
5.29
0.63
GEI 2013
SW11
8.41
5.34
0.63
GEI 2013
SW11
8.41
5.38
0.64
GEI 2013
SW2-1
6.60
12.83
1.95
GEI 2013
SW2-1
6.60
14.80
2.24
GEI 2013
SW2-1
6.60
20.13
3.05
GEI 2013
SW2-1
6.60
26.75
4.06
GEI 2013
SW2-1
6.60
30.48
4.62
GEI 2013
SW2-1
6.60
12.51
1.90
GEI 2013
SW2-1
6.60
16.70
2.53
GEI 2013
SW2-1
6.60
17.21
2.61
GEI 2013
SW2-1
6.60
18.27
2.77
GEI 2013
SW2-1
6.60
20.66
3.13
GEI 2013
SW2-1
9.14
13.31
1.46
GEI 2013
SW2-1
9.14
15.63
1.71
GEI 2013
SW2-1
9.14
15.77
1.73
GEI 2013
SW2-1
9.14
16.79
1.84
GEI 2013
SW2-1
9.14
17.00
1.86
GEI 2013
SW2-1
9.14
18.21
1.99
GEI 2013
SW2-1
9.14
19.39
2.12
GEI 2013
SW2-1
9.14
22.50
2.46
GEI 2013
SW1
7.82
9.11
1.16
GEI 2013
SW1
7.82
9.15
1.17
GEI 2013
SW1
7.82
11.15
1.43
GEI 2013
SW1
7.82
11.23
1.44
GEI 2013
SW1
7.82
13.76
1.76
GEI 2013
SW1
7.82
9.82
1.26
GEI 2013
SW1
7.82
8.45
1.08
GEI 2013
SW1
7.82
8.88
1.14
GEI 2013
SW1
7.82
9.41
1.20
B-lll

-------
Fathead minnow (Pimephales promelas)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
SW1
7.82
11.07
1.42
GEI2013
SW1
6.54
7.01
1.07
GEI 2013
SW1
6.54
7.86
1.20
GEI 2013
SW1
6.54
7.98
1.22
GEI 2013
SW1
6.54
8.23
1.26
GEI 2013
SW1
6.54
8.50
1.30
GEI 2013
SW1
6.54
9.48
1.45
GEI 2013
SW1
6.54
9.95
1.52
GEI 2013
SW1
6.54
10.09
1.54
GEI 2013
SW1
6.54
10.19
1.56
GEI 2013
SW4-1
3.33
5.88
1.77
GEI 2013
SW4-1
3.33
5.89
1.77
GEI 2013
SW4-1
3.33
6.07
1.83
GEI 2013
SW4-1
3.33
6.61
1.99
GEI 2013
SW4-1
3.33
6.87
2.07
GEI 2013
SW4-1
3.33
4.85
1.46
GEI 2013
SW4-1
3.33
5.25
1.58
GEI 2013
SW4-1
3.33
5.39
1.62
GEI 2013
SW4-1
3.33
6.11
1.84
GEI 2013
SW4-1
3.33
6.67
2.01
GEI 2013
SW9
4.45
5.57
1.25
GEI 2013
SW9
4.45
5.93
1.33
GEI 2013
SW9
4.45
6.14
1.38
GEI 2013
SW9
4.45
6.20
1.39
GEI 2013
SW9
4.45
6.56
1.47
GEI 2013
SW9
4.45
7.57
1.70
GEI 2013
SW88
3.96
4.73
1.20
GEI 2013
SW88
3.96
4.96
1.25
GEI 2013
SW88
3.96
5.55
1.40
GEI 2013
SW88
3.96
5.56
1.41
GEI 2013
SW88
3.96
6.32
1.60
GEI 2013
SW88
3.96
5.13
1.30
GEI 2013
SW88
3.96
5.86
1.48
GEI 2013
SW88
3.96
6.07
1.53
GEI 2013
CC1
3.76
3.79
1.01
GEI 2013
CC1
3.76
5.23
1.39
GEI 2013
CC1
3.76
7.36
1.96
GEI 2013
CC1
3.76
8.69
2.31
GEI 2013
CC1
3.76
9.07
2.41
GEI 2013
CC1
4.69
5.92
1.26
GEI 2013
CC1
4.69
7.68
1.64
GEI 2013
CC1
4.69
7.59
1.62
B-112

-------
Fathead minnow (Pimephales promelas)
Study
Site
r
^invert
Cfish
Ratio
GEI2013
CC1
4.69
6.49
1.39
GEI2013
CC1
4.69
7.14
1.52
GEI 2013
CC1
5.86
6.68
1.14
GEI 2013
CC1
5.86
7.73
1.32
GEI 2013
CC1
5.86
7.88
1.35
GEI 2013
CC1
5.86
8.45
1.44
GEI 2013
CC1
5.86
11.69
2.00
GEI 2013
CC1
5.86
9.21
1.57
GEI 2013
CC1
5.86
9.70
1.66
GEI 2013
LG1
3.56
4.81
1.35
GEI 2013
LG1
3.56
4.86
1.37
GEI 2013
LG1
3.56
5.05
1.42
GEI 2013
LG1
3.56
5.47
1.54
GEI 2013
LG1
3.56
5.56
1.56
GEI 2013
LG1
3.56
3.72
1.05
GEI 2013
LG1
3.56
4.09
1.15
GEI 2013
LG1
3.56
3.26
0.92
GEI 2013
LG1
3.56
3.35
0.94
GEI 2013
LG1
3.56
4.20
1.18
GEI 2013
LG1
3.39
3.60
1.06
GEI 2013
LG1
3.39
3.89
1.15
GEI 2013
LG1
3.39
4.27
1.26
GEI 2013
LG1
3.39
4.45
1.31
GEI 2013
LG1
3.39
5.18
1.53
GEI 2013
LG1
3.39
5.51
1.63
Grassoetal. 1995
17
1.91
6.59
3.45
Grassoetal. 1995
17
1.91
6.60
3.46
Grassoetal. 1995
17
1.91
7.30
3.82
Grassoetal. 1995
10
1.85
2.74
1.48
Grassoetal. 1995
10
1.85
2.79
1.51
Grassoetal. 1995
10
1.85
2.90
1.57
Lambing et al. 1994
S46
6.20
5.10
0.82
Lambing et al. 1994
S48
3.05
2.50
0.82
Lambing et al. 1994
Sll
14.50
11.00
0.76
Lambing et al. 1994
Sll
14.50
33.00
2.28
Lambing et al. 1994
S34
14.00
25.00
1.79
Lambing et al. 1994
S39
5.85
7.90
1.35
Lambing et al. 1994
S39
5.85
21.00
3.59
Lemly 1985
Badin Lake
5.70
1.50
0.26
Lemly 1985
Belews Lake
51.15
13.60
0.27
Lemly 1985
High Rock Lake
9.05
1.89
0.21
GEI 2014
DC-4
19.42
27.69
1.43
B-113

-------
Fathead minnow (Pimephales promelas)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
DC-4
19.42
27.88
1.44
GEI2014
DC-4
19.42
23.05
1.19
GEI 2014
DC-4
19.42
30.61
1.58
GEI 2014
DC-4
18.10
19.32
1.07
GEI 2014
DC-4
18.10
14.48
0.80
GEI 2014
DC-4
18.10
25.42
1.40
GEI 2014
FOC-2
2.18
4.13
1.90
GEI 2014
FOC-2
2.18
4.10
1.89
GEI 2014
FOC-2
2.18
5.50
2.53
GEI 2014
FOC-2
2.18
4.85
2.23
GEI 2014
FOC-2
2.18
4.65
2.14
GEI 2014
SC-1
5.85
7.95
1.36
GEI 2014
SC-1
5.85
7.62
1.30
GEI 2014
SC-1
5.85
7.88
1.35
GEI 2014
SC-1
5.85
8.46
1.45
GEI 2014
SC-1
5.85
7.29
1.25
GEI 2014
SC-1
5.85
8.94
1.53
GEI 2014
SC-1
5.85
8.28
1.42
GEI 2014
SC-1
5.85
8.62
1.47
GEI 2014
SC-1
5.85
6.17
1.05
GEI 2014
SC-1
5.85
6.09
1.04
GEI 2014
SC-1
5.85
9.23
1.58
GEI 2014
SC-1
5.85
10.00
1.71
GEI 2014
SC-1
5.85
8.12
1.39
GEI 2014
SC-1
5.85
6.71
1.15
GEI 2014
SC-1
5.85
8.34
1.43
GEI 2014
SC-1
4.94
10.22
2.07
GEI 2014
SC-1
4.94
10.86
2.20
GEI 2014
SC-1
4.94
9.82
1.99
GEI 2014
SC-1
4.94
9.45
1.91
GEI 2014
SC-1
4.94
10.30
2.09
GEI 2014
SC-2
14.33
15.86
1.11
GEI 2014
SC-2
14.33
15.08
1.05
GEI 2014
SC-2
14.33
13.33
0.93
GEI 2014
SC-2
14.33
12.04
0.84
GEI 2014
SC-2
14.33
13.82
0.96
GEI 2014
SC-2
11.44
9.64
0.84
GEI 2014
SC-2
11.44
14.94
1.31
GEI 2014
SC-2
11.44
10.91
0.95
GEI 2014
SC-2
11.44
16.06
1.40
GEI 2014
SC-2
11.44
14.60
1.28
GEI 2014
SC-2
12.75
13.81
1.08
B-114

-------
Fathead minnow (Pimephales promelas)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
SC-2
12.75
14.10
1.11
GEI2014
SC-2
12.75
10.68
0.84
GEI 2014
SC-3
11.41
11.65
1.02
GEI 2014
SC-3
11.41
10.95
0.96
GEI 2014
SC-3
11.41
10.84
0.95
GEI 2014
SC-3
11.41
13.48
1.18
GEI 2014
SC-3
8.58
7.70
0.90
GEI 2014
SC-3
8.58
6.46
0.75
GEI 2014
SC-3
8.58
6.97
0.81
GEI 2014
SC-3
8.58
10.64
1.24
GEI 2014
SC-3
8.58
7.85
0.91
GEI 2014
SC-3
5.75
13.75
2.39
GEI 2014
SC-3
5.75
11.19
1.95
GEI 2014
SC-3
5.75
12.68
2.20
GEI 2014
SC-4
7.39
6.33
0.86
GEI 2014
SC-4
7.39
14.39
1.95
GEI 2014
SC-4
5.18
2.72
0.53
GEI 2014
SC-4
5.18
11.95
2.31
GEI 2014
SC-4
5.18
7.98
1.54
GEI 2014
SC-4
5.18
6.75
1.30
GEI 2014
SC-6
39.87
72.33
1.81
GEI 2014
SC-6
39.87
76.00
1.91
GEI 2014
SC-6
39.87
64.73
1.62
GEI 2014
SC-6
39.87
54.09
1.36
GEI 2014
SC-6
39.87
64.64
1.62
GEI 2014
SC-6
34.35
76.89
2.24
GEI 2014
SC-6
34.35
89.67
2.61
GEI 2014
SC-6
34.35
63.32
1.84
GEI 2014
SC-6
34.35
88.44
2.57
GEI 2014
SC-6
34.35
44.15
1.29
GEI 2014
SC-6
27.54
51.65
1.88
GEI 2014
SC-6
27.54
35.92
1.30
GEI 2014
SC-6
27.54
25.55
0.93
GEI 2014
SC-6
27.54
54.25
1.97
GEI 2014
SC-6
27.54
48.94
1.78
GEI 2014
SC-6
27.54
204.26
7.42
GEI 2014
SC-6
27.54
143.62
5.22
GEI 2014
SC-6
27.54
192.93
7.01
GEI 2014
SC-6
27.54
171.89
6.24
GEI 2014
SC-6
27.54
171.36
6.22
GEI 2014
SC-8
22.62
43.12
1.91
GEI 2014
SC-8
22.62
43.36
1.92
B-115

-------
Fathead minnow (Pimephales promelas)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
SC-8
22.62
55.81
2.47
GEI2014
SC-8
22.62
44.60
1.97
GEI 2014
SC-8
22.62
41.67
1.84
GEI 2014
SC-9
30.36
37.18
1.22
GEI 2014
SC-9
30.36
41.32
1.36
GEI 2014
SC-9
30.36
37.20
1.23
GEI 2014
SC-9
30.36
33.25
1.10
GEI 2014
SC-9
30.36
41.94
1.38
GEI 2014
SC-8
21.77
99.40
4.57
GEI 2014
SC-8
21.77
54.47
2.50
GEI 2014
SC-8
21.77
59.07
2.71
GEI 2014
SC-8
21.77
43.70
2.01
GEI 2014
SC-8
21.77
50.18
2.31
GEI 2014
SC-9
25.06
40.82
1.63
GEI 2014
SC-9
25.06
61.80
2.47
GEI 2014
SC-9
25.06
44.74
1.79
GEI 2014
SC-9
25.06
52.97
2.11
GEI 2014
SC-8
14.15
52.46
3.71
GEI 2014
SC-8
14.15
29.43
2.08
GEI 2014
SC-8
14.15
44.58
3.15
GEI 2014
SC-8
14.15
33.44
2.36
GEI 2014
SC-8
14.15
42.86
3.03
GEI 2014
SC-8
14.15
128.33
9.07
GEI 2014
SC-8
14.15
173.33
12.25
GEI 2014
SC-8
14.15
132.34
9.35
GEI 2014
SC-8
14.15
124.90
8.83
GEI 2014
SC-8
14.15
177.97
12.58
Q
O
i O
o

W	frfl
('
Median ratio: 1.57
R
F
df
P
0.35
185.7
344
<0.001
B-116

-------
Flannelmouth sucker (Catostomus latipinnis)
Study

Site
c
^invert
Cfish
Ratio
Butler et al.
1991
10
4.80
2.50
0.52
Butler et al.
1991
7
29.80
22.00
0.74
Butler et al.
1991
4
3.90
1.70
0.44
Butler et al.
1991
9
4.10
1.50
0.37
Butler et al.
1991
9
4.10
6.00
1.46
Butler et al.
1993
PI
1.95
1.50
0.77
Butler et al.
1993
PI
1.50
2.40
1.60
Butler et al.
1993
LP3
1.12
0.92
0.83
Butler et al.
1993
LP3
1.12
1.40
1.26
Butler et al.
1993
LP4
3.20
2.40
0.75
Butler et al.
1993
LP4
3.20
2.60
0.81
Butler et al.
1994
CRC
7.50
12.00
1.60
Butler et al.
1994
LZA1
19.00
17.00
0.89
Butler et al.
1994
BSW1
5.00
9.60
1.92
Butler et al.
1994
C0L1
1.50
1.90
1.27
Butler et al.
1994
COL1
1.50
0.50
0.33
Butler et al.
1994
COL1
1.50
0.60
0.40
Butler et al.
1994
COL1
1.50
0.63
0.42
Butler et al.
1994
COL1
1.50
0.92
0.61
Butler et al.
1994
COL1
1.50
1.00
0.67
Butler et al.
1994
COL1
1.50
1.60
1.07
Butler et al.
1994
COL1
1.50
1.70
1.13
Butler et al.
1994
COL1
1.50
1.80
1.20
Butler et al.
1994
COL1
1.50
1.90
1.27
Butler et al.
1994
RB3
1.60
29.00
18.13
Butler et al.
1994
RBI
21.00
4.60
0.22
Butler et al.
1994
LSW1
3.90
6.70
1.72
Butler et al.
1994
PSW1
3.70
9.40
2.54
Butler et al.
1995
AK
0.78
1.10
1.42
Butler et al.
1995
AK
0.78
0.90
1.16
Butler et al.
1995
AK
0.78
0.82
1.06
Butler et al.
1995
AK
0.78
1.10
1.42
Butler et al.
1995
HD1
0.83
2.90
3.52
Butler et al.
1995
HD2
0.98
0.49
0.50
Butler et al.
1995
HD2
0.98
0.54
0.55
Butler et al.
1995
HD2
0.98
0.62
0.63
Butler et al.
1995
HD2
0.98
0.96
0.98
Butler et al.
1995
ME2
1.25
1.60
1.28
Butler et al.
1995
ME2
1.25
1.40
1.12
Butler et al.
1995
ME2
1.25
2.00
1.60
Butler et al.
1995
ME2
1.25
2.20
1.76
Butler et al.
1995
ME4
1.55
1.50
0.97
B-117

-------
Flannelmouth sucker (Catostomus latipinnis)
Study

Site
c
^invert
Cfish
Ratio
Butler et al.
1995
ME4
1.55
1.30
0.84
Butler et al.
1995
ME4
1.55
1.90
1.23
Butler et al.
1995
ME4
1.55
2.40
1.55
Butler et al.
1995
ME4
1.55
3.00
1.94
Butler et al.
1995
ME3
2.55
1.70
0.67
Butler et al.
1995
ME3
2.55
1.70
0.67
Butler et al.
1995
ME3
2.55
2.10
0.82
Butler et al.
1995
ME3
2.55
2.40
0.94
Butler et al.
1995
ME3
2.55
3.60
1.41
Butler et al.
1995
SJ1
2.50
1.71
0.68
Butler et al.
1995
SJ1
2.50
1.50
0.60
Butler et al.
1995
SJ1
2.50
2.20
0.88
Butler et al.
1995
SJ1
2.50
0.61
0.24
Butler et al.
1995
SJ1
2.50
1.10
0.44
Butler et al.
1995
SJ1
2.50
4.20
1.68
Butler et al.
1995
YJ2
1.65
1.60
0.97
Butler et al.
1995
YJ2
1.65
2.40
1.45
Butler et al.
1995
MN1
2.70
6.50
2.41
Butler et al.
1995
MN1
2.70
1.70
0.63
Butler et al.
1995
MN1
2.70
4.80
1.78
Butler et al.
1995
MP
1.60
1.20
0.75
Butler et al.
1995
MP
1.60
1.40
0.88
Butler et al.
1997
MN3
2.70
2.30
0.85
Butler et al.
1997
MN3
2.70
2.60
0.96
Butler et al.
1997
MUD
2.30
4.10
1.78
Butler et al.
1997
MUD
2.30
2.70
1.17
Butler et al.
1997
NW2
11.40
11.00
0.96
Butler et al.
1997
MN4
2.65
5.10
1.92
Butler et al.
1997
MN4
2.65
9.60
3.62
Butler et al.
1997
MN5
8.60
8.40
0.98
Butler et al.
1997
MNQ
1.80
2.10
1.17
Butler et al.
1997
MNQ
1.80
3.20
1.78
Butler et al.
1997
MNQ
1.80
3.50
1.94
B-118

-------
Flannelmouth sucker (Catostomus latipinnis)
Study Site
c
^invert
Cfish
Ratio
3u  Q
25 -
O
Median ratio:
0.98

15 -
U1 ooo c
"
R2:
F:
df:
P:
0.36
41.6
73
<0.001

J --	-	
0 5 10 15 20 25 30 35



Gizzard shad (Dorosoma cepedianum)
Study
Site
Mueller etal. 1991
Mueller etal. 1991
Mueller etal. 1991
R2
R1
R1
Cin
Ratio
6.40
8.70
8.70
14.30
7.50
11.00
2.23
0.86
1.26
20
15
Cflsh 10
5
0
Median ratio: 1.26
R
F
df
P
0.74
2.78
1
0.39
2 4 6
^invert.
10
Not used because P > 0.05 and negative
slope.
Goldeye (Hiodon alosoides)
Study Site
r
^invert
Cfish
Ratio
Roddy etal. 1991 18
3.10
2.00
0.65
Roddy etal. 1991 18
3.10
2.10
0.68
Roddy etal. 1991 18
3.10
2.20
0.71
Roddy etal. 1991 18
3.10
2.30
0.74
Roddy etal. 1991 18
3.10
2.40
0.77
Roddy etal. 1991 18
3.10
2.70
0.87
B-119

-------
Goldeye (Hiodon alosoides)
Study

Site
c
^invert
Cfish
Ratio
Roddy et al.
1991
18
3.10
2.90
0.94
Roddy et al.
1991
18
3.10
3.40
1.10
Roddy et al.
1991
18
3.10
3.60
1.16
Roddy et al.
1991
18
3.10
4.70
1.52
5 -|

o



4 -





fish
8
i
2
I
4
Median ratio: 0.82
R2:
0.0
F:
0.0
df:
8
P:
1.0
Not used because no slope and P>0.05.
--invert
Green sunfish (Lepomis cyanettus)
Study

Site
Cjnvert
Cfish
Ratio
Butler et al.
1991
10
4.80
7.90
1.65
Butler et al.
1991
7
29.80
15.20
0.51
Butler et al.
1991
7
29.80
25.10
0.84
Butler et al.
1991
3
6.20
6.40
1.03
Butler et al.
1994
LZA1
19.00
37.00
1.95
Butler et al.
1995
HD1
0.83
1.30
1.58
Butler et al.
1995
HD1
0.83
1.30
1.58
Butler et al.
1995
ME3
2.55
5.00
1.96
Butler et al.
1995
MP
1.60
1.90
1.19
Butler et al.
1997
CHI
7.50
9.50
1.27
Butler et al.
1997
MUD2
3.45
7.60
2.20
Butler et al.
1997
MUD2
3.45
7.00
2.03
Butler et al.
1997
TR25
1.80
4.40
2.44
Butler et al.
1997
TRH
1.60
3.30
2.06
GEI2013

SWA1
2.81
2.96
1.06
GEI2013

SWA1
2.81
3.21
1.14
GEI 2013

SWA1
2.81
3.24
1.16
GEI 2013

SWA1
2.81
3.69
1.32
GEI 2013

SWA1
2.81
3.88
1.38
GEI 2013

SWB
7.44
11.94
1.61
GEI 2013

SW11
8.41
4.54
0.54
B-120

-------
Green sunfish (Lepomis cyanelius)
Study
Site
r
^invert
Cfish
Ratio
GEI2013
SW11
8.41
4.84
0.58
GEI2013
SW11
8.41
5.34
0.63
GEI 2013
SW11
8.41
7.00
0.83
GEI 2013
SW11
8.41
7.13
0.85
GEI 2013
SW9
4.45
4.38
0.98
GEI 2013
SW9
4.45
5.06
1.14
GEI 2013
SW9
4.45
5.53
1.24
GEI 2013
SW9
4.45
5.80
1.30
GEI 2013
SW9
4.45
7.29
1.64
GEI 2013
SW88
3.96
7.14
1.81
GEI 2013
SW88
3.96
7.41
1.87
GEI 2013
LG1
3.39
4.11
1.21
GEI 2013
LG1
3.39
4.33
1.28
GEI 2013
LG1
3.39
5.71
1.68
Roddy etal. 1991
18
3.10
2.80
0.90
Roddy etal. 1991
18
3.10
3.80
1.23
Roddy etal. 1991
18
3.10
4.00
1.29
Roddy etal. 1991
18
3.10
5.20
1.68
Roddy etal. 1991
18
3.10
5.70
1.84
Lemly 1985
Badin Lake
5.70
2.18
0.38
Lemly 1985
Belews Lake
51.15
13.99
0.27
Lemly 1985
High Rock Lake
9.05
2.10
0.23
GEI 2014
C-BCR2
6.81
9.47
1.39
GEI 2014
C-BCR2
6.81
9.29
1.37
GEI 2014
C-BCR2
6.81
8.04
1.18
GEI 2014
C-CC1
4.40
7.23
1.64
GEI 2014
C-CC1
4.40
11.76
2.67
GEI 2014
C-CC2
5.56
4.59
0.83
GEI 2014
C-CC2
5.56
3.04
0.55
GEI 2014
C-CC2
5.56
5.34
0.96
GEI 2014
C-CF1
3.39
3.62
1.07
GEI 2014
C-CF1
3.39
2.95
0.87
GEI 2014
C-CF1
3.39
3.23
0.95
GEI 2014
C-CF1
3.39
5.55
1.64
GEI 2014
C-CLF1
9.30
3.99
0.43
GEI 2014
C-CLF1
9.30
5.23
0.56
GEI 2014
C-CLF1
9.30
4.75
0.51
GEI 2014
C-CLF1
9.30
4.87
0.52
GEI 2014
C-CLF1
9.30
3.58
0.39
GEI 2014
C-CLF2
6.85
5.76
0.84
GEI 2014
C-CLF2
6.85
5.89
0.86
GEI 2014
C-CLF2
6.85
4.78
0.70
B-121

-------
Green sunfish (Lepomis cyanelius)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
C-CLF2
6.85
5.11
0.75
GEI2014
C-CLF2
6.85
4.10
0.60
GEI 2014
C-LFWOC1
9.32
10.10
1.08
GEI 2014
C-TF1
20.00
4.15
0.21
GEI 2014
C-TF1
20.00
3.47
0.17
GEI 2014
C-TF1
20.00
4.14
0.21
GEI 2014
C-TF1
20.00
4.11
0.21
GEI 2014
C-TF1
20.00
3.41
0.17
GEI 2014
C-WOC1
6.65
12.05
1.81
GEI 2014
C-WOC1
6.65
9.60
1.44
GEI 2014
C-WOC1
6.65
8.66
1.30
GEI 2014
C-WOC1
6.65
5.81
0.87
GEI 2014
C-WOC1
6.65
7.54
1.13
GEI 2014
H-BB1
16.29
9.55
0.59
GEI 2014
H-BB1
16.29
18.27
1.12
GEI 2014
H-BB1
16.29
7.08
0.43
GEI 2014
H-BHC1
5.08
3.69
0.73
GEI 2014
H-BHC1
5.08
2.48
0.49
GEI 2014
H-BHC1
5.08
3.29
0.65
GEI 2014
H-BHC1
5.08
3.49
0.69
GEI 2014
H-BHC1
5.08
3.70
0.73
GEI 2014
H-JSB1
4.03
4.83
1.20
GEI 2014
H-JSB1
4.03
2.57
0.64
GEI 2014
H-JSB1
4.03
3.73
0.93
GEI 2014
H-LF1
9.09
3.12
0.34
GEI 2014
H-LF1
9.09
5.96
0.66
GEI 2014
H-LF1
9.09
4.30
0.47
GEI 2014
H-LF1
9.09
4.02
0.44
GEI 2014
H-LF1
9.09
5.29
0.58
GEI 2014
H-MR2
2.14
9.57
4.47
GEI 2014
H-MR2
2.14
5.55
2.59
GEI 2014
H-MR2
2.14
5.80
2.71
GEI 2014
H-MR2
2.14
5.55
2.59
GEI 2014
H-MR2
2.14
6.88
3.22
GEI 2014
H-MR3
3.86
8.09
2.10
GEI 2014
H-MR3
3.86
16.98
4.40
GEI 2014
H-MR3
3.86
6.80
1.76
GEI 2014
H-MR3
3.86
8.52
2.21
GEI 2014
H-MR3
3.86
6.62
1.72
GEI 2014
H-MR4
9.26
9.01
0.97
GEI 2014
H-MR4
9.26
8.78
0.95
GEI 2014
H-MR4
9.26
18.33
1.98
B-122

-------
Green sunfish (Lepomis cyanelius)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
H-MR4
9.26
9.84
1.06
GEI2014
H-MR4
9.26
5.94
0.64
GEI 2014
H-MR5
3.58
5.50
1.54
GEI 2014
H-MR5
3.58
3.52
0.98
GEI 2014
H-MR5
3.58
2.41
0.67
GEI 2014
H-MR5
3.58
3.09
0.86
GEI 2014
H-MR5
3.58
1.94
0.54
GEI 2014
H-MR6
2.49
36.20
14.54
GEI 2014
H-MR6
2.49
2.58
1.04
GEI 2014
H-MR6
2.49
1.94
0.78
GEI 2014
H-MR6
2.49
1.74
0.70
GEI 2014
H-MR6
2.49
2.69
1.08
GEI 2014
H-SB1
10.62
11.90
1.12
GEI 2014
H-SB1
10.62
13.39
1.26
GEI 2014
H-SB1
10.62
7.64
0.72
GEI 2014
H-SB1
10.62
13.45
1.27
GEI 2014
H-SF1
21.05
13.59
0.65
GEI 2014
H-SF1
21.05
14.22
0.68
GEI 2014
H-SF1
21.05
15.27
0.73
GEI 2014
H-SF1
21.05
14.58
0.69
GEI 2014
H-SF1
21.05
11.25
0.53
GEI 2014
H-SF2
13.95
17.74
1.27
GEI 2014
H-SF2
13.95
12.86
0.92
GEI 2014
H-SF2
13.95
12.76
0.91
GEI 2014
H-SF2
13.95
13.41
0.96
GEI 2014
H-SF2
13.95
28.23
2.02
GEI 2014
H-UB1
3.02
4.43
1.47
GEI 2014
H-UB1
3.02
4.91
1.63
GEI 2014
H-UB1
3.02
3.73
1.23
GEI 2014
H-UB1
3.02
8.00
2.65
GEI 2014
H-UB1
3.02
8.36
2.77
GEI 2014
SC-1-25
4.00
12.78
3.19
GEI 2014
SC-2
14.33
9.31
0.65
GEI 2014
SC-2
14.33
7.59
0.53
GEI 2014
SC-2
11.44
10.23
0.89
GEI 2014
SC-2-27
23.76
30.00
1.26
GEI 2014
SC-3
5.75
8.66
1.51
GEI 2014
SC-3
5.75
11.72
2.04
GEI 2014
SC-3
5.75
9.59
1.67
GEI 2014
SC-4
5.18
19.86
3.83
GEI 2014
SC-6
39.87
49.55
1.24
GEI 2014
SC-6
39.87
49.56
1.24
B-123

-------
Green sunfish (Lepomis cyanellus)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
SC-6
39.87
33.71
0.85
GEI2014
SC-6
39.87
28.73
0.72
GEI 2014
SC-6
39.87
40.64
1.02
GEI 2014
SC-6
34.35
41.51
1.21
GEI 2014
SC-6
34.35
48.02
1.40
GEI 2014
SC-6
34.35
68.66
2.00
GEI 2014
SC-6
34.35
41.27
1.20
GEI 2014
SC-6
34.35
48.76
1.42
GEI 2014
SC-6
27.54
28.17
1.02
GEI 2014
SC-6
27.54
106.88
3.88
GEI 2014
SC-6
27.54
114.55
4.16
GEI 2014
SC-6
27.54
100.41
3.65
GEI 2014
SC-6
27.54
153.64
5.58
GEI 2014
SC-6
27.54
145.62
5.29
GEI 2014
S-SC1
11.01
7.47
0.68
ISO -
1*1
U'l
8
Median ratio:
1.12

1'lfl
8
R2:
0.37
as* -,n

F:
93.17
on

df:
160
!lJ O
 	rf	
P:
<0.001
j3>
:o 30
Iowa darter (Etheostoma exile)
Study

Site
c
^invert
Cfish
Ratio
1978
7
3.75
2.10
0.56
1978
20
11.20
36.30
3.24
1978
22
11.30
23.00
2.04
1978
23
15.50
41.90
2.70
B-124

-------
Iowa darter (Etheostoma exile)
Median ratio
P: 0.055
Not used because P > 0.05
invert.
Largemouth bass (Micropterus salmoides)
Study
Site
Cjnvert
Cfish
Ratio
Butler etal. 1995
MP
Cienega de Santa
1.60
1.40
0.88
Garcia-Hernandez et al. 2000
Clara Wetland
3.00
5.10
1.70
GEI2013
SWA1
2.81
3.17
1.13
GEI2013
SW11
8.41
5.02
0.60
GEI 2013
SW11
8.41
5.77
0.69
GEI 2013
SW11
8.41
5.19
0.62
GEI 2013
SW11
8.41
6.26
0.74
GEI 2013
SW11
8.41
6.48
0.77
GEI 2013
SW11
8.41
7.22
0.86
GEI 2013
SW4-1
3.33
5.53
1.66
GEI 2013
SW4-1
3.33
5.65
1.70
GEI 2013
SW4-1
3.33
5.72
1.72
GEI 2013
SW4-1
3.33
5.80
1.74
GEI 2013
SW4-1
3.33
6.34
1.91
GEI 2013
SW4-1
3.33
7.14
2.15
GEI 2013
SW9
4.45
5.78
1.30
GEI 2013
SW9
4.45
5.79
1.30
GEI 2013
SW9
4.45
6.19
1.39
GEI 2013
SW9
4.45
6.87
1.54
GEI 2013
SW9
4.45
7.27
1.63
GEI 2013
SW9
4.45
7.36
1.65
GEI 2013
SW88
3.96
4.87
1.23
GEI 2013
SW88
3.96
5.73
1.45
GEI 2013
SW88
3.96
5.77
1.46
GEI 2013
SW88
3.96
5.93
1.50
GEI 2013
SW88
3.96
6.62
1.67
GEI 2013
SW88
3.96
6.84
1.73
B-125

-------
Largemouth bass (Micropterus salmoides)
Study
Site
r
^invert
Cfish
Ratio
GEI2013
LG1
3.39
4.29
1.27
Saikietal. 1993
ET6
0.85
1.00
1.18
Saikietal. 1993
ET6
0.85
1.40
1.66
Saikietal. 1993
GT5
4.90
6.80
1.39
Saikietal. 1993
GT5
4.90
6.90
1.41
Saikietal. 1993
GT4
4.05
4.70
1.16
Saikietal. 1993
GT4
4.05
4.00
0.99
Saikietal. 1993
SJR2
3.30
2.20
0.67
Saikietal. 1993
SJR2
3.30
2.40
0.73
Saikietal. 1993
SJR3
1.50
1.80
1.20
Saikietal. 1993
SJR3
1.50
1.70
1.13
Saikietal. 1993
SJR1
0.95
0.80
0.85
Saikietal. 1993
SJR1
0.95
1.80
1.90
Saikietal. 1993
ET7
0.86
0.86
1.00
Saikietal. 1993
ET7
0.86
1.00
1.16
Rinellaand Schuler 1992
S. Malheur Lake
1.20
0.92
0.77
Crutchfield 2000
transect 3
11.95
12.52
1.05
Crutchfield 2000
transect 3
11.40
16.67
1.46
Crutchfield 2000
transect 3
9.25
6.83
0.74
Crutchfield 2000
transect 3
9.25
6.99
0.76
Crutchfield 2000
transect 3
8.60
6.59
0.77
Crutchfield 2000
transect 3
8.60
5.69
0.66
Crutchfield 2000
transect 4
20.90
15.53
0.74
Crutchfield 2000
transect 4
20.90
19.68
0.94
Crutchfield 2000
transect 4
15.70
18.95
1.21
Crutchfield 2000
transect 4
15.70
9.43
0.60
Crutchfield 2000
transect 4
16.45
6.83
0.42
Crutchfield 2000
transect 4
18.25
9.43
0.52
GEI 2014
ARB
11.21
20.41
1.82
GEI 2014
ARB
11.21
32.75
2.92
GEI 2014
ARB
11.21
32.73
2.92
GEI 2014
ARB
11.21
29.23
2.61
GEI 2014
ARB
11.21
21.26
1.90
GEI 2014
ARE
20.40
25.35
1.24
GEI 2014
ARE
20.40
20.80
1.02
GEI 2014
ARE
20.40
22.67
1.11
GEI 2014
ARE
20.40
21.57
1.06
GEI 2014
ARE
20.40
16.05
0.79
GEI 2014
ARM
8.51
13.62
1.60
GEI 2014
ARM
8.51
10.13
1.19
GEI 2014
ARM
8.51
12.00
1.41
GEI 2014
ARM
8.51
11.40
1.34
B-126

-------
Largemouth bass (Micropterus salmoides)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
ARM
8.51
8.71
1.02
GEI2014
ARM
7.68
12.52
1.63
GEI 2014
ARM
7.68
13.59
1.77
GEI 2014
ARM
7.68
17.99
2.34
GEI 2014
ARM
7.68
13.49
1.76
GEI 2014
ARM
7.68
15.98
2.08
GEI 2014
ARN
8.06
11.16
1.39
GEI 2014
ARN
8.06
12.89
1.60
GEI 2014
ARN
8.06
16.74
2.08
GEI 2014
ARN
8.06
10.12
1.26
GEI 2014
ARN
8.06
20.08
2.49
GEI 2014
ARN
7.49
17.47
2.33
GEI 2014
ARN
7.49
13.48
1.80
GEI 2014
ARN
7.49
19.44
2.60
GEI 2014
ARN
7.49
21.87
2.92
GEI 2014
ARN
7.49
21.91
2.92
GEI 2014
ARN
7.44
30.73
4.13
GEI 2014
ARN
7.44
40.71
5.47
GEI 2014
ARN
7.44
38.75
5.21
GEI 2014
ARN
7.44
38.24
5.14
GEI 2014
SC-5
15.13
38.13
2.52
GEI 2014
SC-5
15.13
42.86
2.83
GEI 2014
SC-9
30.36
41.87
1.38
GEI 2014
SC-9
25.06
37.84
1.51


1
o
o
o

o


.,-6"
8 
"o >
,,0'"
0)
o

o o
o
ID 15 20
Median ratio: 1.39
R
F
df
P
0.40
61.45
91
<0.001
B-127

-------
Longnose dace (Rhinichthys cataractae)
Study
Site
c
^invert
Cfish
Ratio
Lambing et al. 1994
S33
2.40
5.30
2.21
Mueller etal. 1991
Al
2.70
2.10
0.78
GEI 2013
LG1
3.37
5.05
1.50
GEI 2013
LG1
3.37
5.57
1.65
GEI 2013
LG1
3.37
6.57
1.95
GEI 2013
LG1
3.37
6.75
2.00
GEI 2013
LG1
3.37
10.08
2.99
GEI 2013
LG1
3.39
10.69
3.15
GEI 2013
LG1
3.39
12.77
3.77
GEI 2013
LG1
3.56
8.95
2.52
GEI 2013
LG1
3.56
9.63
2.71
GEI 2013
LG1
3.56
11.41
3.21
GEI 2013
LG1
3.56
11.94
3.36
GEI 2013
LG1
Left Fork White Oak
3.56
12.04
3.39
GEI 2014
Creek, C-LFWOC1
Left Fork White Oak
9.32
8.32
0.89
GEI 2014
Creek, C-LFWOC1
Left Fork White Oak
9.32
9.47
1.02
GEI 2014
Creek, C-LFWOC1
Left Fork White Oak
9.32
7.94
0.85
GEI 2014
Creek, C-LFWOC1
Left Fork White Oak
9.32
6.67
0.72
GEI 2014
Creek, C-LFWOC1
9.32
7.19
0.77
Mueller etal. 1991
T1
5.40
16.90
3.13
Hamilton and Buhl 2005
CC
6.70
13.40
2.00
Hamilton and Buhl 2005
BGS
10.80
10.90
1.01
18
16
14
12
10
8
6
4
2
0
o 8
o
o
o
o
-e-
Median ratio: 2.00
R2:
0.01
F:
0.19
df:
20
P:
0.83
Not used because P > 0.05
10
12
B-128

-------
Longnose sucker (Catostomus catostomus)
Study
Site
c
^invert
Cfish
Ratio
Minnow 2007
FL17
3.03
1.40
0.46
Butler etal. 1994
NFK2
3.10
2.10
0.68
Butler etal. 1994
NFK2
3.10
2.50
0.81
Butler etal. 1994
NFK2
3.10
2.70
0.87
Butler etal. 1994
NFK2
3.10
2.80
0.90
Butler etal. 1994
NFK2
3.10
2.90
0.94
Butler etal. 1994
NFK2
3.10
3.00
0.97
Butler etal. 1994
NFK2
3.10
3.20
1.03
Butler etal. 1994
NFK2
3.10
3.30
1.06
Butler etal. 1994
NFK2
3.10
3.40
1.10
Butler etal. 1994
NFK2
3.10
4.00
1.29
Mueller etal. 1991
T1
5.40
3.60
0.67
Minnow 2007
FL17
21.22
7.90
0.37
Median ratio: 0.90
Mottled sculpin (Cottus bairdii)
Study
Site
c
^invert
Cfish
Ratio
Hamilton and Buhl 2004
use
0.50
5.30
10.60
Butler etal. 1993
LP2
1.00
2.20
2.20
Butler etal. 1993
LP2
1.00
3.10
3.10
Butler etal. 1993
LP3
1.12
3.90
3.50
Butler etal. 1993
LP3
1.12
4.20
3.77
Butler etal. 1993
LP3
1.12
4.90
4.39
Butler etal. 1993
PI
1.50
5.10
3.40
Butler etal. 1993
PI
1.50
6.40
4.27
Butler etal. 1993
PI
1.50
6.70
4.47
Hamilton and Buhl 2004
ShpC
1.90
4.10
2.16
Butler etal. 1993
PI
1.95
7.30
3.74
Butler etal. 1994
NFK3
2.00
5.80
2.90
B-129

-------
Mottled sculpin (Cottus bairdii)
Study
Site
c
^invert
Cflsh
Ratio
Butler etal. 1997
MN2
2.20
2.60
1.18
Butler etal. 1997
CHK
2.40
3.10
1.29
Butler etal. 1997
CHK
2.40
4.40
1.83
Lambing et al. 1994
S33
2.40
3.70
1.54
Butler etal. 1991
12
2.80
4.20
1.50
Butler etal. 1997
MN1
2.90
3.20
1.10
Butler etal. 1997
MN1
2.90
3.40
1.17
Butler etal. 1994
NFK2
3.10
6.40
2.06
Butler etal. 1991
4
3.90
2.60
0.67
Butler etal. 1991
4
3.90
4.40
1.13
Butler etal. 1991
10
4.80
5.00
1.04
Hamilton and Buhl 2005
UAC
5.00
6.20
1.24
Butler etal. 1991
3
6.20
6.50
1.05
Hamilton and Buhl 2004
ACM
6.70
8.30
1.24
Hamilton and Buhl 2005
CC
6.70
8.20
1.22
Butler etal. 1993
F2
7.50
9.90
1.32
Hamilton and Buhl 2004
LBR
7.70
5.20
0.68
Hamilton and Buhl 2005
DC
8.70
12.00
1.38
Hamilton and Buhl 2005
BGS
10.80
12.30
1.14
Hamilton and Buhl 2004
DVC
12.80
8.80
0.69
Butler etal. 1994
HCC1
21.00
5.60
0.27
Median ratio: 1.38
Mountain whitefish (Prosopium williamsoni)
Study	Site	Cinvert	Cflsh	Ratio
Low and Mullins 1990	7	1.60	1.40	0.88
McDonald and Strosher 1998	ER745	2.74	4.17	1.52
Minnow 2007	EL12	4.01	6.60	1.65
McDonald and Strosher 1998	ER747	4.29	4.93	1.15
B-130

-------
Mountain whitefish (Prosopium williamsoni)
Study

Site
r
^invert

Cfish
Ratio
Minnow 2007
MI3
6.21

9.12
1.47
Minnow 2007
MI2
6.69

10.16
1.52
Minnow 2007
ELI
7.08

9.12
1.29
Minnow 2007
F023
10.00

10.20
1.02
14 i






12 -






10 -
o .
o$*
/ o




8 "


Median ratio:
1.38

Tish g .






4 -



R2:
0.83

2 -
O


F:
30.27

A



df:
6

u





0 5
10 15

P:
<0.001


c
v invert.





Northern pike (Esox lucius)
Study

Site
Cjnvert

Cfish
Ratio
Butler et al.
1995
TT
1.07

2.96
2.78
Butler et al.
1995
TT
1.07

3.24
3.04
Butler et al.
1995
TT
1.07

1.65
1.55
Butler et al.
1995
TT
1.07

1.18
1.11
Butler et al.
1995
TT
1.07

1.90
1.78
Butler et al.
1995
TT
1.07

1.80
1.69
Butler et al.
1995
PU
0.61

0.93
1.52
Butler et al.
1995
PU
0.61

1.40
2.30
Muscatello et al. 2008
David Lake
1.39

0.78
0.56
Muscatello et al. 2008
Delta Lake
9.38

17.02
1.81
Muscatello et al. 2008
Unknown Lake
15.71

28.28
1.80
Muscatello and Janz 2009
Indigo Lake
0.36

0.75
2.08
Muscatello and Janz 2009
Vulture Lake
1.62

1.26
0.78
B-131

-------
Northern pike (Esox lucius)
25 -

Median ratio:
1.78
20 -

R2:
0.99
15 '

F:
982.9
10

df:
11


P:
<0.001
10
Northern plains killfish (Fun(lulus kansae)
Study
Site
c
^invert
Cfish
Ratio
Birkner 1978
3
3.10
7.70
2.48
Birkner 1978
11
5.65
5.00
0.88
Birkner 1978
23
15.50
23.10
1.49
Birkner 1978
27
34.60
31.90
0.92
Birkner 1978
30
45.05
57.40
1.27
70
60
50
40
-fish 30
20
10
0
I	1	1	1	1
10 20 30 40 50
Median ratio: 1.27
R2:
0.93
F:
37.8
df:
3
P:
0.008
-'invert.
B-132

-------
Rainbow trout (Oncorhynchus my kiss)
Study

Site
r
^invert
Cfish
Ratio
Butler et al.
991
4
3.90
3.50
0.90
Butler et al.
993
F2
4.80
7.60
1.58
Butler et al.
993
F2
3.90
7.60
1.95
Butler et al.
993
LP2
1.00
0.78
0.78
Butler et al.
993
LP2
1.00
1.40
1.40
Butler et al.
993
LP3
1.12
1.90
1.70
Butler et al.
994
GUN2
28.00
5.40
0.19
Butler et al.
994
HCC1
21.00
16.48
0.78
Butler et al.
994
NFK3
2.00
4.70
2.35
Butler et al.
994
NFK2
3.10
21.98
7.09
Butler et al.
994
NFK2
3.10
3.60
1.16
Butler et al.
995
MP
1.60
1.88
1.18
Butler et al.
995
MP
1.60
2.30
1.44
Butler et al.
995
MP
1.60
2.50
1.56
Butler et al.
995
MP
1.60
2.10
1.31
Butler et al.
997
CHK
2.40
1.41
0.59
Butler et al.
997
CHK
2.40
2.20
0.92
Butler et al.
997
CHK
2.40
2.50
1.04
Butler et al.
997
CHK
2.40
2.80
1.17
Butler et al.
997
CHK
2.40
2.90
1.21
Butler et al.
997
MN3
2.70
2.28
0.84
Butler et al.
997
MN3
2.70
2.60
0.96
Butler et al.
997
MN3
2.70
4.90
1.81
Butler et al.
997
MN2
2.20
2.10
0.95
Butler et al.
997
MN2
2.20
2.80
1.27
Butler et al.
997
MN1
2.90
2.50
0.86
Butler et al.
997
MN1
2.90
2.60
0.90
Butler et al.
997
MN1
2.90
3.20
1.10
Butler et al.
997
WBR
5.05
5.10
1.01
Low and Mu
lins 1990
5
5.60
2.60
0.46
Casey 2005

Deerlick Creek
4.45
1.29
0.29
Casey 2005

Deerlick Creek
4.45
4.15
0.93
Casey 2005

Luscar Creek
9.95
6.88
0.69
Casey 2005

Luscar Creek
9.95
17.04
1.71
B-133

-------
Rainbow trout (Oncorhynchus mykiss)

o
0

Median ratio:
1.07



R2:
0.19



F:
7.52
J3'"

o
df:
32



P:
0.002
10 !5 20
Invert*
Red shiner (Cyprinella lutrensis)
Study
Site
r
^invert
Cfish
Ratio
Butler etal. 1991
3
6.20
7.70
1.24
Butler etal. 1994
IW
8.35
83.00
9.94
Butler etal. 1994
AD
2.70
7.30
2.70
Butler etal. 1994
LW
3.00
19.00
6.33
Butler etal. 1994
LSW1
3.90
14.00
3.59
Butler etal. 1995
ME4
1.55
5.10
3.29
Butler etal. 1995
ME3
2.55
4.60
1.80
Butler etal. 1995
ME3
2.55
4.20
1.65
Butler etal. 1995
SJ1
2.50
3.50
1.40
Butler etal. 1995
YJ2
1.65
4.50
2.73
Butler etal. 1997
MN4
2.65
4.20
1.58
Butler etal. 1997
MN5
8.60
4.40
0.51
May et al. 2008
KR
17.20
7.03
0.41
May et al. 2008
NSCL
10.70
7.36
0.69
May et al. 2008
NSCU
10.50
7.24
0.69
May et al. 2008
NSK
8.81
5.81
0.66
May et al. 2008
NSP
24.00
8.62
0.36
May et al. 2008
SSAL
11.50
9.00
0.78
May et al. 2008
SSAU
8.35
11.20
1.34
May et al. 2008
sso
10.00
7.16
0.72
May et al. 2008
ssw
7.60
10.00
1.32
Mueller etal. 1991
A3
6.00
8.10
1.35
Mueller etal. 1991
A2
8.50
7.90
0.93
Lemly 1985
Badin Lake
5.70
2.10
0.37
Lemly 1985
Belews Lake
51.15
18.25
0.36
Lemly 1985
High Rock Lake
9.05
2.18
0.24
GEI2014
ARE
20.40
49.84
2.44
B-134

-------
Red shiner (Cyprinelia lutrensis)
Study
Site
r
^invert
Cfish
Ratio
GEI2014
ARE
20.40
80.50
3.95
GEI2014
ARE
20.40
20.57
1.01
GEI 2014
ARE
20.40
33.37
1.64
GEI 2014
ARE
20.40
26.50
1.30
GEI 2014
ARN
7.44
27.50
3.70
GEI 2014
ARN
7.44
23.58
3.17
GEI 2014
ARN
7.44
21.74
2.92
GEI 2014
FC-4
18.65
21.20
1.14
GEI 2014
FC-4
18.65
32.68
1.75
GEI 2014
FC-4
18.65
25.73
1.38
GEI 2014
GC-1
9.33
10.78
1.16
GEI 2014
GC-1
9.33
9.97
1.07
GEI 2014
SC-2
12.75
10.44
0.82
GEI 2014
SC-2
12.75
10.87
0.85
GEI 2014
SC-3
5.75
12.05
2.10
GEI 2014
SC-3
5.75
12.17
2.12
GEI 2014
SC-3
5.75
9.93
1.73
GEI 2014
SC-3
5.75
9.93
1.73
GEI 2014
SC-4
7.39
12.26
1.66
GEI 2014
SC-4
7.39
11.68
1.58
GEI 2014
SC-4
7.39
14.15
1.92
GEI 2014
SC-4
5.18
9.58
1.85
GEI 2014
SC-4
5.18
8.43
1.63
GEI 2014
SC-4
5.18
10.83
2.09
GEI 2014
SC-5
15.13
17.96
1.19
GEI 2014
SC-5
15.13
34.71
2.29
GEI 2014
SC-5
15.13
34.05
2.25
GEI 2014
SC-5
15.13
37.28
2.46
GEI 2014
SC-5
15.13
32.18
2.13
GEI 2014
SC-6
39.87
53.60
1.34
GEI 2014
SC-6
39.87
37.00
0.93
GEI 2014
SC-6
39.87
35.11
0.88
GEI 2014
SC-6
39.87
51.39
1.29
GEI 2014
SC-6
39.87
42.31
1.06
GEI 2014
SC-8
22.62
29.20
1.29
GEI 2014
SC-9
25.06
22.55
0.90
GEI 2014
SC-9
25.06
18.02
0.72
GEI 2014
SC-9
25.06
25.94
1.04
GEI 2014
SC-9
25.06
18.68
0.75
GEI 2014
SC-9
25.06
14.28
0.57
GEI 2014
SC-9
25.06
31.67
1.26
GEI 2014
SC-9
25.06
20.43
0.82
B-135

-------
Red shiner (Cyprinetta lutrensis)
Study
Site
<8

cin
^flsh
Ratio
GEI2014

SC-9

25.06
22.27
0.89
GEI2014

SC-9

25.06
27.05
1.08
GEI 2014

SC-9

25.06
25.28
1.01

9(1
ft. 
-o
o

Median ratio:
1.31


SO -
i  
30  8

o
If"
8
--S
R2:
F:
df:
P:
0.28
26.57
70
<0.001

Redside shiner (Ri ch ardson ius bait eat us)
Study
Site
c
^invert
Cfish
Ratio
Hamilton and Buhl 2004
ACM
6.70
6.00
0.90
Hamilton and Buhl 2004
LBR
7.70
2.70
0.35
Hamilton and Buhl 2005
BGS
10.80
13.20
1.22
GEI 2014
Bond Creek, BC-3
3.02
3.58
1.19
GEI 2014
Bond Creek, BC-3
3.02
3.44
1.14
GEI 2014
Bond Creek, BC-3
3.02
3.44
1.14
GEI 2014
Bond Creek, BC-3
3.02
4.64
1.54
GEI 2014
Bond Creek, BC-3
3.02
3.26
1.08
GEI 2014
Bond Creek, BC-3
5.87
3.18
0.54
GEI 2014
Bond Creek, BC-3
5.87
3.37
0.57
GEI 2014
Bond Creek, BC-3
5.87
2.79
0.47
GEI 2014
Bond Creek, BC-3
3.02
3.58
1.19
B-136

-------
Redside shiner (Ri ch ardson ius bait eat us)
14
12
10
o

o
10
12
Median ratio:
1.08
R2:
0.47
F:
7.84
df:
9
P:
0.011
( .
Roundtail chub (Gila robusta)
Study

Site
Cjnvert
Cfish
Ratio
Butler et al.
1994
COL1
1.50
2.20
1.47
Butler et al.
1994
COL1
1.50
2.50
1.67
Butler et al.
1994
COL1
1.50
2.70
1.80
Butler et al.
1994
COL1
1.50
3.30
2.20
Butler et al.
1994
COL1
1.50
3.70
2.47
Butler et al.
1994
COL1
1.50
4.10
2.73
Butler et al.
1994
COL1
1.50
5.10
3.40
Butler et al.
1994
COL1
1.50
5.30
3.53
Butler et al.
1994
COL1
1.50
26.00
17.33
Butler et al.
1994
RB3
1.60
5.40
3.38
Butler et al.
1995
MP
1.60
4.20
2.63
Butler et al.
1997
MUD
2.30
4.60
2.00
Butler et al.
1994
AD
2.70
7.10
2.63
Butler et al.
1994
LW
3.00
4.50
1.50
Butler et al.
1994
NFK2
3.10
6.10
1.97
Butler et al.
1994
PSW1
3.70
7.70
2.08
Butler et al.
1994
LSW1
3.90
5.80
1.49
Butler et al.
1991
10
4.80
1.90
0.40
Butler et al.
1994
TGC
4.90
10.00
2.04
Butler et al.
1994
BSW1
5.00
8.10
1.62
Butler et al.
1993
F2
7.50
7.30
0.97
Butler et al.
1994
CRC
7.50
19.00
2.53
Butler et al.
1994
IW
8.35
8.50
1.02
Butler et al.
1997
NW2
11.40
6.90
0.61
Butler et al.
1994
RBI
21.00
5.90
0.28
Butler et al.
1994
GUN2
28.00
6.80
0.24
B-137

-------
Roundtail chub (Gila robusta)
30
25
20
Cfisi, 15
10
5
0

8
mF~
0 o

10
c.
Median ratio: 1.98
R
F
df
P
0.01
0.18
24
0.834
Not used because P > 0.05
20
30
invert.
Sand shiner (Notropis stramineus)
Study
Site
r
^invert
Cfish
Ratio
GEI2013
SW1
6.54
8.43
1.29
GEI2013
SW1
6.54
9.02
1.38
GEI 2013
SW1
6.54
9.66
1.48
GEI 2013
SW1
6.54
11.21
1.71
GEI 2013
SW1
6.54
11.85
1.81
GEI 2013
SW1
6.54
11.94
1.83
GEI 2013
SW1
6.54
13.50
2.06
GEI 2013
SW1
6.54
14.05
2.15
GEI 2013
SW1
6.54
14.14
2.16
GEI 2013
SW2-1
6.60
18.70
2.84
GEI 2013
SW2-1
6.60
19.33
2.93
GEI 2013
SW2-1
6.60
19.77
3.00
GEI 2013
SW2-1
6.60
20.39
3.09
GEI 2013
SW2-1
6.60
23.70
3.59
GEI 2013
SWB
7.06
8.27
1.17
GEI 2013
SWB
7.06
9.01
1.28
GEI 2013
SWB
7.06
9.81
1.39
GEI 2013
SWB
7.06
10.22
1.45
GEI 2013
SW1
7.82
11.33
1.45
GEI 2013
SW1
7.82
12.05
1.54
GEI 2013
SW1
7.82
12.22
1.56
GEI 2013
SW1
7.82
12.55
1.60
GEI 2013
SW1
7.82
12.65
1.62
GEI 2013
SW1
7.82
12.68
1.62
GEI 2013
SW1
7.82
14.13
1.81
GEI 2013
SW1
7.82
14.43
1.85
GEI 2013
SW1
7.82
15.87
2.03
B-138

-------
Sand shiner (Natr apis stramineus)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
SW1
7.82
16.63
2.13
GEI2013
SW2-1
9.14
17.84
1.95
GEI 2013
SW2-1
9.14
18.21
1.99
GEI 2013
SW2-1
9.14
18.98
2.08
GEI 2013
SW2-1
9.14
20.12
2.20
GEI 2013
SW2-1
9.14
20.73
2.27
GEI 2014
Arkansas River, ARE
20.40
21.50
1.05
GEI 2014
Arkansas River, ARE
20.40
23.20
1.14
GEI 2014
Arkansas River, ARE
20.40
22.64
1.11
GEI 2014
Arkansas River, ARE
20.40
25.24
1.24
GEI 2014
Arkansas River, ARE
20.40
29.70
1.46
GEI 2014
Arkansas River, ARM
8.51
10.67
1.25
GEI 2014
Arkansas River, ARN
8.06
9.69
1.20
GEI 2014
Arkansas River, ARN
8.06
9.27
1.15
GEI 2014
Arkansas River, ARN
8.06
9.96
1.24
GEI 2014
Arkansas River, ARN
8.06
9.29
1.15
GEI 2014
Arkansas River, ARN
8.06
8.86
1.10
GEI 2014
Arkansas River, ARN
7.49
13.60
1.82
GEI 2014
Arkansas River, ARN
7.49
18.34
2.45
GEI 2014
Arkansas River, ARN
7.49
16.46
2.20
GEI 2014
Arkansas River, ARN
7.49
19.64
2.62
GEI 2014
Fountain Creek, FC-4
14.59
13.95
0.96
GEI 2014
Fountain Creek, FC-4
14.59
9.34
0.64
GEI 2014
Fountain Creek, FC-4
14.59
14.06
0.96
GEI 2014
Fountain Creek, FC-4
14.59
28.26
1.94
GEI 2014
Fountain Creek, FC-4
14.59
10.53
0.72
GEI 2014
Fountain Creek, FC-4
17.15
10.28
0.60
GEI 2014
Fountain Creek, FC-4
17.15
23.76
1.39
GEI 2014
Fountain Creek, FC-4
17.15
14.77
0.86
GEI 2014
Fountain Creek, FC-4
17.15
23.13
1.35
GEI 2014
Fountain Creek, FC-4
17.15
25.62
1.49
GEI 2014
Fountain Creek, FCP
6.13
17.62
2.88
GEI 2014
Fountain Creek, FCP
6.13
7.32
1.19
GEI 2014
Fountain Creek, FCP
6.13
7.14
1.17
GEI 2014
Fountain Creek, FCP
6.13
6.05
0.99
GEI 2014
Fountain Creek, FCP
6.13
7.11
1.16
GEI 2014
Fountain Creek, FCP
6.35
15.93
2.51
GEI 2014
St. Charles River, SC-4
6.29
11.92
1.90
GEI 2014
St. Charles River, SC-4
6.29
15.14
2.41
GEI 2014
St. Charles River, SC-4
6.29
8.94
1.42
GEI 2014
St. Charles River, SC-4
6.29
10.33
1.64
GEI 2014
St. Charles River, SC-4
6.29
11.58
1.84
B-139

-------
Sand shiner (Notropis stramineus)
Study Site
c
^invert
Cfish
Ratio
35
30
25
20
15
10
5
0
10	15
(
vin veils
20
25
Median ratio: 1.56
R
F
df
P
0.32
32.15
67
<0.001
Sculpin (Cottoidea)
Study
Site
Cjnvert
Cfish
Ratio
Mason et al. 2000
BK
1.43
1.16
0.81
Mason et al. 2000
BK
1.43
2.35
1.64
Mason et al. 2000
BK
1.43
2.64
1.84
Formation 2012
SFTC-1
1.63
9.31
5.71
Formation 2012
SFTC-1
2.42
5.68
2.35
Formation 2012
SFTC-1
2.49
5.87
2.36
Formation 2012
CC-75
3.11
5.03
1.62
Formation 2012
CC-75
3.11
5.58
1.79
Formation 2012
CC-350
3.16
6.47
2.05
Formation 2012
CC-350
3.16
7.12
2.26
Formation 2012
SFTC-1
3.21
3.75
1.17
Formation 2012
CC-75
3.97
3.77
0.95
Formation 2012
CC-75
4.16
7.08
1.70
Formation 2012
CC-75
4.16
7.19
1.73
Formation 2012
CC-350
4.20
5.28
1.26
Formation 2012
CC-150
4.46
5.04
1.13
Formation 2012
CC-150
4.46
6.01
1.35
Formation 2012
CC-150
4.70
5.14
1.09
Formation 2012
CC-3A
5.45
11.65
2.14
Formation 2012
CC-3A
5.45
14.45
2.65
Formation 2012
CC-3A
5.48
11.47
2.09
Formation 2012
CC-150
7.03
10.73
1.53
Formation 2012
DC-600
7.83
7.96
1.02
Formation 2012
DC-600
7.83
8.62
1.10
B-140

-------
Sculpin (Cottoidea)
Study
Site
c
^invert
Cfish
Ratio
Formation 2012
DC-600
8.53
7.87
0.92
Formation 2012
DC-600
8.53
8.50
1.00
Formation 2012
DC-600
8.65
7.63
0.88
Formation 2012
LSV-4
9.54
18.28
1.92
Formation 2012
LSV-4
9.54
20.01
2.10
Formation 2012
HS-3
11.40
18.57
1.63
Formation 2012
HS-3
11.40
21.85
1.92
Formation 2012
CC-350
11.45
9.53
0.83
Formation 2012
CC-350
11.45
10.03
0.88
Formation 2012
CC-1A
12.24
8.34
0.68
Formation 2012
CC-1A
12.24
9.94
0.81
Formation 2012
CC-1A
12.24
17.47
1.43
Formation 2012
CC-1A
12.57
7.78
0.62
Formation 2012
HS-3
13.41
26.63
1.99
Formation 2012
CC-1A
13.55
12.63
0.93
Formation 2012
CC-150
14.32
7.35
0.51
Formation 2012
CC-3A
14.50
20.20
1.39
Formation 2012
HS
15.70
23.23
1.48
Formation 2012
HS
15.70
23.25
1.48
Formation 2012
HS
18.70
10.95
0.59
Formation 2012
LSV-2C
22.62
11.38
0.50
Formation 2012
LSV-2C
22.62
17.47
0.77
Formation 2012
HS-3
24.70
23.93
0.97
Formation 2012
LSV-2C
26.31
18.85
0.72
Formation 2012
HS-3
26.55
23.68
0.89
Formation 2012
LSV-2C
26.95
20.32
0.75
Formation 2012
HS
27.80
35.93
1.29
Formation 2012
HS
27.80
41.30
1.49
Formation 2012
LSV-2C
30.00
25.95
0.87
Median ratio: 1.29
B-141

-------
Shorthead redhorse (Moxostoma macrolepidotum)
Study

Site
Cjnvert
Cfish
Ratio
Roddy et al.
1991
18
3.10
2.80
0.90
Roddy et al.
1991
18
3.10
2.90
0.94
Roddy et al.
1991
18
3.10
2.90
0.94
Roddy et al.
1991
18
3.10
3.10
1.00
Roddy et al.
1991
18
3.10
3.30
1.06
Roddy et al.
1991
18
3.10
3.40
1.10
Roddy et al.
1991
18
3.10
3.50
1.13
Roddy et al.
1991
18
3.10
3.60
1.16
Roddy et al.
1991
18
3.10
3.70
1.19
Roddy et al.
1991
18
3.10
3.80
1.23
Roddy et al.
1991
18
3.10
3.80
1.23
Median ratio: 1.10
R2:
0.00
F:
0.00
df:
9
P:
1.0
0	1	2	3	4 Not used because P > 0.05
(
invert.
Smallmouth bass (Micropterus dolomieu)
Study
Site
Cjnvert
Cfish
Ratio
Butler etal. 1995
SU
1.85
1.55
0.84
Butler etal. 1995
su
1.85
1.22
0.66
Butler etal. 1995
SU
1.85
0.98
0.53
Butler etal. 1995
su
1.85
1.14
0.62
Butler etal. 1995
SU
1.85
1.50
0.81
Butler etal. 1995
su
1.85
1.50
0.81
Butler etal. 1995
MP
1.60
1.90
1.19
Butler etal. 1997
MNP3
6.15
12.00
1.95
Mueller etal. 1991
R1
8.70
2.90
0.33
Mueller etal. 1991
R1
8.70
4.10
0.47
GEI2014
ARE
20.40
24.61
1.21
GEI2014
ARE
20.40
22.97
1.13
GEI 2014
ARE
20.40
13.28
0.65
B-142

-------
Smallmouth bass (Micropterus dolomieu)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
ARE
20.40
19.06
0.93
GEI2014
ARE
20.40
19.11
0.94
GEI 2014
ARM
7.10
8.25
1.16
GEI 2014
ARM
7.10
8.04
1.13
GEI 2014
ARM
7.10
7.72
1.09
GEI 2014
ARM
7.10
6.21
0.87
GEI 2014
ARM
7.10
6.51
0.92
GEI 2014
C-SC1
11.30
8.94
0.79
GEI 2014
C-SC1
11.30
7.68
0.68
Median ratio:	0.86
R2:	0.84
F:	107.1
df:	20
P:	<0.001
Speckled dace (Rhinichthys osculus)
Study

Site
Cjnvert
Cfish
Ratio
Butler et al.
1991
10
4.80
4.80
1.00
Butler et al.
1991
9
4.10
5.70
1.39
Butler et al.
1991
3
6.20
6.50
1.05
Butler et al.
1993
SP2
2.75
12.00
4.36
Butler et al.
1993
B2
1.35
5.80
4.30
Butler et al.
1993
B1
1.25
4.40
3.52
Butler et al.
1993
B1
1.25
4.40
3.52
Butler et al.
1993
D1
1.20
3.50
2.92
Butler et al.
1993
D1
1.20
3.70
3.08
Butler et al.
1993
D1
1.20
3.40
2.83
Butler et al.
1993
D2
1.45
4.90
3.38
Butler et al.
1993
D2
1.45
6.80
4.69
Butler et al.
1993
D2
1.45
6.50
4.48
Butler et al.
1993
F2
3.90
8.90
2.28
Butler et al.
1993
PI
1.95
5.50
2.82
Butler et al.
1993
SP1
2.95
7.30
2.47
Cn 15
5 \	^ 
8
0
o
o

%
0	5	10	15	20	25
B-143

-------
Speckled dace (Rhinichthys osculus)
Study

Site
r
^invert
Cfish
Ratio
Butler et al.
1993
SP1
2.95
8.90
3.02
Butler et al.
1993
SP1
2.95
7.00
2.37
Butler et al.
1993
U1
2.45
3.60
1.47
Butler et al.
1993
U1
2.45
6.90
2.82
Butler et al.
1993
U1
2.45
7.30
2.98
Butler et al.
1993
U1
2.45
9.20
3.76
Butler et al.
1993
U1
2.45
9.40
3.84
Butler et al.
1993
U1
2.45
9.80
4.00
Butler et al.
1993
LP3
1.12
6.00
5.38
Butler et al.
1993
LP4
3.20
8.70
2.72
Butler et al.
1993
R2
4.30
17.10
3.98
Butler et al.
1993
R2
3.90
6.00
1.54
Butler et al.
1993
ST2
4.50
15.70
3.49
Butler et al.
1993
ST2
4.10
8.50
2.07
Butler et al.
1993
ST2
4.10
10.70
2.61
Butler et al.
1993
ST2
3.35
9.30
2.78
Butler et al.
1993
R1
4.00
8.50
2.13
Butler et al.
1993
ST1
2.25
6.80
3.02
Butler et al.
1993
SB2
3.60
12.10
3.36
Butler et al.
1993
SB2
3.75
7.80
2.08
Butler et al.
1993
SB2
3.75
10.80
2.88
Butler et al.
1993
SB1
2.15
10.00
4.65
Butler et al.
1993
SB1
2.15
9.50
4.42
Butler et al.
1993
SB1
2.15
7.80
3.63
Butler et al.
1993
WSB2
4.75
15.60
3.28
Butler et al.
1993
WSB2
3.60
11.70
3.25
Butler et al.
1993
WSB2
3.00
6.20
2.07
Butler et al.
1993
WSB2
3.00
7.60
2.53
Butler et al.
1994
CRC
7.50
13.00
1.73
Butler et al.
1994
CF1
3.60
6.10
1.69
Butler et al.
1994
GUN2
28.00
8.90
0.32
Butler et al.
1994
IW
8.35
10.00
1.20
Butler et al.
1994
LZA1
19.00
28.00
1.47
Butler et al.
1994
NFK3
2.00
7.10
3.55
Butler et al.
1994
NFK2
3.10
6.90
2.23
Butler et al.
1994
NFK2
3.10
4.80
1.55
Butler et al.
1994
NFK2
3.10
5.40
1.74
Butler et al.
1994
NFK2
3.10
5.70
1.84
Butler et al.
1994
NFK2
3.10
6.10
1.97
Butler et al.
1994
NFK2
3.10
6.20
2.00
Butler et al.
1994
NFK2
3.10
6.30
2.03
Butler et al.
1994
NFK2
3.10
6.40
2.06
B-144

-------
Speckled dace (Rhinichthys osculus)
Study

Site
r
^invert
Cfish
Ratio
Butler et al.
1994
NFK2
3.10
6.70
2.16
Butler et al.
1994
NFK2
3.10
7.40
2.39
Butler et al.
1994
NFK2
3.10
8.70
2.81
Butler et al.
1994
TGC
4.90
12.00
2.45
Butler et al.
1994
BSW1
5.00
15.00
3.00
Butler et al.
1994
C0L1
1.50
2.30
1.53
Butler et al.
1994
COL1
1.50
5.00
3.33
Butler et al.
1994
COL1
1.50
7.30
4.87
Butler et al.
1994
COL1
1.50
7.40
4.93
Butler et al.
1994
COL1
1.50
8.40
5.60
Butler et al.
1994
COL1
1.50
8.60
5.73
Butler et al.
1994
COL1
1.50
9.30
6.20
Butler et al.
1994
COL1
1.50
9.60
6.40
Butler et al.
1994
COL1
1.50
11.00
7.33
Butler et al.
1994
RB3
1.60
93.00
58.13
Butler et al.
1994
SMF
4.80
7.80
1.63
Butler et al.
1994
LW
3.00
62.00
20.67
Butler et al.
1994
LSW1
3.90
83.00
21.28
Butler et al.
1994
PSW1
3.70
13.00
3.51
Butler et al.
1995
AK
0.78
4.30
5.55
Butler et al.
1995
AK
0.78
3.10
4.00
Butler et al.
1995
AK
0.78
4.00
5.16
Butler et al.
1995
DD
0.86
5.60
6.51
Butler et al.
1995
DD
0.86
4.40
5.12
Butler et al.
1995
DD
0.86
6.00
6.98
Butler et al.
1995
HD1
0.83
2.80
3.39
Butler et al.
1995
HD1
0.83
3.20
3.88
Butler et al.
1995
HD1
0.83
5.30
6.42
Butler et al.
1995
ME1
3.40
6.40
1.88
Butler et al.
1995
ME2
1.25
6.10
4.88
Butler et al.
1995
ME3
2.55
2.80
1.10
Butler et al.
1995
ME3
2.55
7.00
2.75
Butler et al.
1995
ME3
2.55
5.50
2.16
Butler et al.
1995
NW
5.10
8.70
1.71
Butler et al.
1995
SJ1
2.50
4.30
1.72
Butler et al.
1995
SJ1
2.50
5.10
2.04
Butler et al.
1995
SJ1
2.50
2.90
1.16
Butler et al.
1995
YJ2
1.65
6.50
3.94
Butler et al.
1995
YJ2
1.65
6.30
3.82
Butler et al.
1995
YJ2
1.65
7.10
4.30
Butler et al.
1995
MN1
2.70
5.50
2.04
Butler et al.
1997
CHK
2.40
5.20
2.17
B-145

-------
Speckled dace (Rhinichthys osculus)
Study

Site
c
^invert
Cflsh
Ratio
Butler et al.
1997
CHK
2.40
3.80
1.58
Butler et al.
1997
MN3
2.70
6.00
2.22
Butler et al.
1997
MN3
2.70
4.30
1.59
Butler et al.
1997
MN2
2.20
2.70
1.23
Butler et al.
1997
MN2
2.20
3.60
1.64
Butler et al.
1997
MN1
2.90
3.70
1.28
Butler et al.
1997
MUD
2.30
7.20
3.13
Butler et al.
1997
MUD
2.30
6.10
2.65
Butler et al.
1997
NW2
11.40
11.00
0.96
Butler et al.
1997
WBR
5.05
9.70
1.92
Butler et al.
1997
WBR
5.05
5.50
1.09
Butler et al.
1997
MN4
2.65
7.90
2.98
Butler et al.
1997
MN5
8.60
14.00
1.63
Butler et al.
1997
MNQ
1.80
5.90
3.28
Hamilton and Buhl 2004
DVC
12.80
7.50
0.59
Hamilton and Buhl 2004
use
0.50
6.90
13.80
Hamilton and Buhl 2004
ACM
6.70
8.50
1.27
Hamilton and Buhl 2004
LBR
7.70
5.60
0.73
Hamilton and Buhl 2005
LiB
5.40
5.80
1.07
Hamilton and Buhl 2005
SLC
9.70
15.20
1.57
70 

Median ratio:
2.76
(W 



SfJ

R2:
0.01
40 -

F:
1.76
30 
O
df:
118
_
oPQ 
		
P:
0.177
Not used because P > 0.05
Sucker (Catostomidae)
Study	Site	Cinvert	Cflsh	Ratio
Butler etal. 1995	HD2	0.98	0.68	0.69
Butler etal. 1995	HD2	0.98	0.76	0.78
Butler etal. 1993	D1	1.20	2.30	1.92
Rinella and Schuler 1992	Malheur Lake	1.20	1.60	1.33
B-146

-------
Sucker (Catostomidae)
Study
Site
c
^invert
Cfish
Ratio
Butler etal. 1993
B2
1.35
1.80
1.33
Butler etal. 1995
YJ2
1.65
2.20
1.33
Butler etal. 1993
PI
1.95
1.50
0.77
Butler etal. 1993
U1
2.45
2.30
0.94
Butler etal. 1993
U1
2.45
3.60
1.47
Butler etal. 1991
12
2.80
2.10
0.75
Butler etal. 1994
NFK2
3.10
35.00
11.29
Butler etal. 1993
SB2
3.60
5.10
1.42
Butler etal. 1993
R2
3.90
5.00
1.28
Butler etal. 1993
R2
4.30
2.20
0.51
Butler etal. 1993
ST2
4.50
10.00
2.22
Butler etal. 1993
WSB2
4.75
11.80
2.48
Butler etal. 1993
F2
7.50
4.20
0.56
Wish
O so
Median ratio: 1.33
R
F
df
P
0.07
1.10
15
0.360
Not used because P > 0.05
-'Invert.
Sunfish (Centrarchidae)
Study
Site
Cjnvert
Cfish
Ratio
Welsh and Maughan 1994
outfall drain
1.15
2.30
2.00
Welsh and Maughan 1994
Pretty Water
1.16
1.80
1.56
Welsh and Maughan 1994
Hart Mine Marsh
1.20
2.40
2.00
Welsh and Maughan 1994
outfall drain
1.30
2.10
1.62
Welsh and Maughan 1994
outfall drain
1.30
2.80
2.15
Welsh and Maughan 1994
Pretty Water
1.50
1.60
1.07
Welsh and Maughan 1994
Old Channel
1.50
2.00
1.33
Welsh and Maughan 1994
Pretty Water
1.50
2.30
1.53
Welsh and Maughan 1994
Cibola Lake
1.85
5.90
3.19
Welsh and Maughan 1994
Cibola Lake
1.90
5.30
2.79
Welsh and Maughan 1994
Cibola Lake
1.90
7.60
4.00
Welsh and Maughan 1994
Oxbow Lake
3.60
11.00
3.06
B-147

-------
Sunfish (Centrarchidae)
9.14
8.10
Median ratio:
2.00
R2:
0.38
F:
6.66
df:
11
P:
0.013
GEI2013
-fish
12 1
10 -
8
6
4
2
0
SW2-1
2 4 6
Cjnvert.
10
0.89
Tui chub (Gila bicolor)
Study

Site
Cjnvert
Cfish
Ratio
Sorenson & Schwarzbach 1991
5
0.49
1.20
2.45
Sorenson & Schwarzbach 1991
4
0.76
1.00
1.32
Rinellaand Schuler 1992
Harney Lake
2.05
3.10
1.51
4 1





3 -





3 -


Median ratio:
1.51

2 -
C
*--fisb 2 -


R2:
0.94

1 -


F:
15.9

1 -


df:
1




P:
0.175

U "1
1 1
2 2 3
Not used because P >
0.05


c
^ invert.




Walleye (Sander vitreus)
Study

Site
Cjnvert
Cfish
Ratio
Butler et al.
1995
XT
1.07
1.86
1.75
Butler et al.
1995
XT
1.07
1.62
1.52
Butler et al.
1995
XT
1.07
1.70
1.60
Butler et al.
1995
XT
1.07
1.70
1.60
Butler et al.
1995
XT
1.07
2.00
1.88
B-148

-------
Walleye (Sander vitreus)
Study

Site

c
^invert
Cfish
Ratio
Butler etal. 1995

XT

1.07
1.60
1.50
Butler etal. 1995

PU

0.61
1.72
2.82
Butler etal. 1995

PU

0.61
1.05
1.73
Butler etal. 1995

PU

0.61
0.81
1.33
Butler etal. 1995

PU

0.61
1.00
1.64
Butler etal. 1995

PU

0.61
0.89
1.46
Mueller etal. 1991

R1

8.70
2.40
0.28
Peterson et al. 1991

7

3.83
4.27
1.11
Peterson et al. 1991

7

3.83
4.79
1.25
Peterson et al. 1991

7

3.83
6.76
1.77
Peterson et al. 1991

7

3.83
8.35
2.18
9 
8
6 -
O
o


Median ratio: 1.60

5 
CfKi, ^
o
o


R2: 0.28

3 



F: 5.46



o
df: 14

2 erf
1 8



P: 0.018

0
0 2
4

8 10










Western niosquitofish (Gambusia affinis)
Study
Site
Cjnvert
Cfish
Ratio
GEI2013
SWA1
3.64
2.91
0.80
GEI2013
SWA1
2.81
3.01
1.07
GEI 2013
SWA1
2.81
3.49
1.24
GEI 2013
SWA1
2.81
3.66
1.30
GEI 2013
SWA1
2.81
3.89
1.39
GEI 2013
SWA1
2.81
4.27
1.52
Lemly 1985
Badin Lake
5.70
3.35
0.59
Lemly 1985
Belews Lake
51.15
27.20
0.53
Lemly 1985
High Rock Lake
9.05
3.54
0.39
Saiki and Lowe 1987
Kesterson Pond 11
60.65
130.00
2.14
Saiki and Lowe 1987
Kesterson Pond 11
60.65
104.00
1.71
Saiki and Lowe 1987
Kesterson Pond 2
177.00
224.00
1.27
Saiki and Lowe 1987
Kesterson Pond 2
177.00
247.00
1.40
Saiki and Lowe 1987
Kesterson Pond 8
102.50
164.00
1.60
B-149

-------
Western niosquitofish (Gambusia affinis)
Study

Site
r
^invert
Cflsh
Ratio
Saiki and Lowe 1987
Kesterson Pond 8
102.50
223.00
2.18
Saiki and Lowe 1987
San Luis Drain
190.00
149.00
0.78
Saiki and Lowe 1987
San Luis Drain
190.00
332.00
1.75
Saiki and Lowe 1987
Volta Pond 26
1.42
1.28
0.90
Saiki and Lowe 1987
Volta Pond 26
1.42
1.24
0.87
Saiki and Lowe 1987
Volta Wasteway
2.23
1.35
0.61
Saiki and Lowe 1987
Volta Wasteway
2.23
1.36
0.61
Saiki et al.
1993
ET6
0.85
1.00
1.18
Saiki et al.
1993
ET6
0.85
1.30
1.54
Saiki et al.
1993
GT5
4.90
16.00
3.27
Saiki et al.
1993
GT5
4.90
11.00
2.24
Saiki et al.
1993
GT4
4.05
4.50
1.11
Saiki et al.
1993
GT4
4.05
4.90
1.21
Saiki et al.
1993
SJR2
3.30
4.50
1.36
Saiki et al.
1993
SJR2
3.30
2.20
0.67
Saiki et al.
1993
SJR3
1.50
1.70
1.13
Saiki et al.
1993
SJR3
1.50
2.00
1.33
Saiki et al.
1993
SJR1
0.95
0.95
1.01
Saiki et al.
1993
SJR1
0.95
1.30
1.38
Saiki et al.
1993
ET7
0.86
0.90
1.05
Saiki et al.
1993
ET7
0.86
1.00
1.16

3%
3
25D
2ms
Jb
o
O
O
50	100
c,
o
Median ratio: 1.21
R
F
df
P
0.89
263.3
33
<0.001
White sucker (Catostomus commersonii)
Study	Site	Cinvert	Cflsh	Ratio
Butler etal. 1993	LP3	1.12	2.50	2.24
Butler etal. 1993	B1	1.25	2.60	2.08
Butler etal. 1993	D2	1.45	1.90	1.31
Butler etal. 1993	D2	1.45	2.50	1.72
B-150

-------
White sucker (Catostomus commersonii)
Study
Site

c
^invert
Cfish
Ratio
Butler etal. 1993
PI

1.50
1.70
1.13
Butler etal. 1993
PI

1.50
1.80
1.20
Butler etal. 1995
MP

1.60
1.40
0.88
Butler etal. 1995
SU

1.85
1.20
0.65
GEI2014
Arkansas River
ARB
11.21
14.90
1.33
GEI2014
Arkansas River
ARB
11.21
20.39
1.82
GEI 2014
Arkansas River
ARB
11.21
13.82
1.23
GEI 2014
Arkansas River
ARB
11.21
8.36
0.75
GEI 2014
Arkansas River
ARB
11.21
10.88
0.97
GEI 2014
Arkansas River
ARB
11.21
21.55
1.92
GEI 2014
Arkansas River
ARB
11.21
18.70
1.67
GEI 2014
Arkansas River
ARB
11.21
24.53
2.19
GEI 2014
Arkansas River
ARB
11.21
15.02
1.34
GEI 2014
Arkansas River
ARB
11.21
28.29
2.52
GEI 2014
Arkansas River
ARE
20.4
18.21
0.89
GEI 2014
Arkansas River
ARE
20.4
19.54
0.96
GEI 2014
Arkansas River
ARE
20.4
15.27
0.75
GEI 2014
Arkansas River
ARE
20.4
11.37
0.56
GEI 2014
Arkansas River
ARE
20.4
17.86
0.88
GEI 2014
Arkansas River
ARE
20.4
10.62
0.52
GEI 2014
Arkansas River
ARE
20.4
17.51
0.86
GEI 2014
Arkansas River
ARE
20.4
24.66
1.21
GEI 2014
Arkansas River
ARE
20.4
18.92
0.93
GEI 2014
Arkansas River
ARE
20.4
21.70
1.06
GEI 2014
Arkansas River
ARM
8.51
10.13
1.19
GEI 2014
Arkansas River
ARM
8.51
9.24
1.09
GEI 2014
Arkansas River
ARM
8.51
8.30
0.97
GEI 2014
Arkansas River
ARM
8.51
10.09
1.19
GEI 2014
Arkansas River
ARM
7.68
16.18
2.11
GEI 2014
Arkansas River
ARM
7.68
13.21
1.72
GEI 2014
Arkansas River
ARM
7.68
11.96
1.56
GEI 2014
Arkansas River
ARM
7.68
8.58
1.12
GEI 2014
Arkansas River
ARM
7.68
9.73
1.27
GEI 2014
Arkansas River
ARM
7.1
8.19
1.15
GEI 2014
Arkansas River
ARM
7.1
7.96
1.12
GEI 2014
Arkansas River
ARN
7.49
9.15
1.22
GEI 2014
Arkansas River
ARN
7.49
8.61
1.15
GEI 2014
Arkansas River
ARN
7.49
7.06
0.94
GEI 2014
Arkansas River
ARN
7.49
11.57
1.54
GEI 2014
Arkansas River
ARN
7.49
11.56
1.54
GEI 2014
Arkansas River
ARN
7.44
21.20
2.85
GEI 2014
Arkansas River
ARN
7.44
23.28
3.13
B-151

-------
White sucker (Catostomus commersonii)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
Arkansas River, ARN
7.44
20.85
2.80
GEI2014
Arkansas River, ARN
7.44
25.91
3.48
GEI 2014
Bond Creek, BC-2
2.96
3.06
1.03
GEI 2014
Bond Creek, BC-2
2.96
3.54
1.19
GEI 2014
Bond Creek, BC-2
2.96
3.04
1.03
GEI 2014
Bond Creek, BC-3
3.02
2.76
0.91
GEI 2014
Bond Creek, BC-3
3.02
2.85
0.94
GEI 2014
Bond Creek, BC-3
3.02
2.47
0.82
GEI 2014
Bond Creek, BC-3
3.02
2.03
0.67
GEI 2014
Bond Creek, BC-3
3.02
2.23
0.74
GEI 2014
Cow Camp Creek, CC-2
5.65
4.46
0.79
GEI 2014
Cow Camp Creek, CC-2
5.65
4.45
0.79
GEI 2014
Cow Camp Creek, CC-2
5.65
6.19
1.09
GEI 2014
Cow Camp Creek, CC-2
5.65
4.71
0.83
GEI 2014
Cow Camp Creek, CC-2
5.65
5.38
0.95
GEI 2014
Seng Creek, C-SC1
11.302
20.32
1.80
GEI 2014
Dry Creek, DC-4
19.42
26.07
1.34
GEI 2014
Dry Creek, DC-4
19.42
22.55
1.16
GEI 2014
Dry Creek, DC-4
19.42
14.29
0.74
GEI 2014
Dry Creek, DC-4
19.42
14.25
0.73
GEI 2014
Dry Creek, DC-4
19.42
14.67
0.76
GEI 2014
Dry Creek, DC-4
18.1
29.83
1.65
GEI 2014
Dry Creek, DC-4
18.1
30.65
1.69
GEI 2014
Dry Creek, DC-4
18.1
20.87
1.15
GEI 2014
Dry Creek, DC-4
18.1
12.06
0.67
GEI 2014
Fountain Creek, FC-4
18.65
24.54
1.32
GEI 2014
Fountain Creek, FCP
6.13
5.33
0.87
GEI 2014
Fountain Creek, FCP
6.13
5.88
0.96
GEI 2014
Fountain Creek, FCP
6.13
5.88
0.96
GEI 2014
Fountain Creek, FCP
6.13
5.75
0.94
GEI 2014
Fountain Creek, FCP
6.13
4.37
0.71
GEI 2014
Fountain Creek, FCP
5.38
8.50
1.58
GEI 2014
Fountain Creek, FCP
5.38
5.94
1.10
GEI 2014
Fountain Creek, FCP
5.38
5.97
1.11
GEI 2014
Fountain Creek, FCP
5.38
5.76
1.07
GEI 2014
Fountain Creek, FCP
5.38
5.61
1.04
GEI 2014
Fountain Creek, FCP
6.35
15.82
2.49
GEI 2014
Fountain Creek, FCP
6.35
5.68
0.90
GEI 2014
Fountain Creek, FCP
6.35
10.17
1.60
GEI 2014
Fountain Creek, FCP
6.35
12.34
1.94
GEI 2014
Fountain Creek, FCP
6.35
10.64
1.68
GEI 2014
Foidel Creek, FOC-2
2.175
1.74
0.80
B-152

-------
White sucker (Catostomus commersonii)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
Foidel Creek, FOC-2
2.175
1.25
0.57
GEI2014
Foidel Creek, FOC-2
2.175
1.76
0.81
GEI 2014
Foidel Creek, FOC-2
2.175
2.11
0.97
GEI 2014
Foidel Creek, FOC-2
2.175
1.64
0.76
GEI 2014
Foidel Creek, FOC-2
2.175
2.11
0.97
GEI 2014
Foidel Creek, FOC-2
2.175
2.29
1.05
GEI 2014
Grassy Creek, GC-2
4.195
4.45
1.06
GEI 2014
Grassy Creek, GC-2
4.195
4.42
1.05
GEI 2014
Grassy Creek, GC-2
4.195
2.51
0.60
GEI 2014
Grassy Creek, GC-3
4.535
2.78
0.61
GEI 2014
Grassy Creek, GC-3
4.535
2.76
0.61
GEI 2014
Grassy Creek, GC-3
4.535
2.84
0.63
GEI 2014
Grassy Creek, GC-3
4.535
4.37
0.96
GEI 2014
Grassy Creek, GC-3
4.535
2.89
0.64
GEI 2014
Grassy Creek, GC-3
4.545
4.29
0.94
GEI 2014
Grassy Creek, GC-3
4.545
3.16
0.69
GEI 2014
Grassy Creek, GC-3
4.545
2.76
0.61
GEI 2014
Grassy Creek, GC-3
4.34
4.12
0.95
GEI 2014
Grassy Creek, GC-3
4.35
3.96
0.91
GEI 2014
Grassy Creek, GC-4
5.1
0.93
0.18
GEI 2014
Grassy Creek, GC-4
5.1
1.40
0.27
GEI 2014
Grassy Creek, GC-4
5.1
4.12
0.81
GEI 2014
Grassy Creek, GC-4
5.76
7.04
1.22
GEI 2014
Grassy Creek, GC-4
5.76
7.42
1.29
GEI 2014
Grassy Creek, GC-4
5.76
4.22
0.73
GEI 2014
Grassy Creek, GC-4
5.76
4.75
0.82
GEI 2014
Grassy Creek, GC-4 US
13.16
6.32
0.48
GEI 2014
Grassy Creek, GC-4 US
13.16
4.74
0.36
GEI 2014
Grassy Creek, GC-4 US
13.16
4.98
0.38
GEI 2014
Grassy Creek, GC-4 US
13.16
4.75
0.36
GEI 2014
Grassy Creek, GC-4 US
13.16
4.88
0.37
GEI 2014
Middle Creek, MC-1
3.21
2.03
0.63
GEI 2014
Middle Creek, MC-1
3.21
2.18
0.68
GEI 2014
Middle Creek, MC-1
3.21
1.79
0.56
GEI 2014
Middle Creek, MC-1
3.21
2.28
0.71
GEI 2014
Middle Creek, MC-1
3.21
2.54
0.79
GEI 2014
Middle Creek, MC-2
4.19
2.58
0.62
GEI 2014
Middle Creek, MC-2
4.19
1.90
0.45
GEI 2014
Middle Creek, MC-2
4.19
1.86
0.44
GEI 2014
Middle Creek, MC-2
4.19
1.90
0.45
GEI 2014
Middle Creek, MC-2
4.19
2.10
0.50
GEI 2014
St. Charles River, SC-3
5.75
7.74
1.35
B-153

-------
White sucker (Catostomus commersonii)
Study
Site
c
^invert
Cfish
Ratio
GEI2014
St. Charles River, SC-3
5.75
12.34
2.15

St. Charles River at US



GEI2014
Hwy 50, SC-5
15.13
46.76
3.09

St. Charles River at US



GEI 2014
Hwy 50, SC-5
15.13
38.23
2.53

St. Charles River at US



GEI 2014
Hwy 50, SC-5
15.13
46.59
3.08

St. Charles River at US



GEI 2014
Hwy 50, SC-5
15.13
60.66
4.01

St. Charles River at US



GEI 2014
Hwy 50, SC-5
15.13
43.70
2.89
GEI 2014
St. Charles River, SC-9
25.06
35.97
1.44
GEI 2014
St. Charles River, SC-9
25.06
38.53
1.54

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
7.62
1.76

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
8.20
1.89

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
6.10
1.41

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
22.44
5.18

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
4.29
0.99

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
4.79
1.11

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
47.18
10.88

St. Charles River at 1-25,



GEI 2014
SC-I
4.335
4.88
1.12
GEI 2014
Wildhorse Creek, WHC
56.14
5.43
0.10
GEI 2014
Wildhorse Creek, WHC
34.24
36.23
1.06
GEI 2014
Wildhorse Creek, WHC
34.24
10.22
0.30
GEI 2014
Wildhorse Creek, WHC
34.24
52.16
1.52
GEI 2014
Wildhorse Creek, WHC
34.24
40.81
1.19
GEI 2014
Wildhorse Creek, WHC
34.24
26.45
0.77
GEI 2014
Wildhorse Creek, WHC
62.34
61.90
0.99
GEI 2014
Wildhorse Creek, WHC
62.34
15.88
0.25
GEI 2014
Wildhorse Creek, WHC
62.34
27.40
0.44
GEI 2014
Wildhorse Creek, WHC
62.34
23.10
0.37
Muscatello and Janz 2009
Indigo Lake
0.36
0.99
2.75
Muscatello and Janz 2009
Vulture Lake
1.62
3.37
2.08
Grassoetal. 1995
17
1.91
2.84
1.49
Grassoetal. 1995
17
1.91
3.19
1.67
Grassoetal. 1995
17
1.91
3.44
1.80
Grassoetal. 1995
17
1.91
3.64
1.91
B-154

-------
White sucker (Catostomus commersonii)
Study
Site
c
^invert
Cfish
Ratio
Grassoetal. 1995
17
1.91
4.00
2.09
Grassoetal. 1995
17
1.91
4.01
2.10
Butler etal. 1994
NFK3
2.00
3.90
1.95
Butler etal. 1993
ST1
2.25
4.90
2.18
Lambing et al. 1994
S33
2.40
3.50
1.46
Mueller etal. 1991
Al
2.70
4.20
1.56
GEI2013
SWA1
2.81
2.83
1.01
GEI2013
SWA1
2.81
3.89
1.39
GEI 2013
SWA1
2.81
4.18
1.49
Mason et al. 2000
HCRT
2.81
0.81
0.29
Mason et al. 2000
HCRT
2.81
1.43
0.51
Mason et al. 2000
HCRT
2.81
1.43
0.51
Butler etal. 1993
WSB2
3.00
3.90
1.30
Butler etal. 1993
SP2
3.15
3.50
1.11
Butler etal. 1993
LP4
3.20
2.80
0.88
GEI 2013
SW4-1
3.33
3.01
0.91
GEI 2013
SW4-1
3.33
3.45
1.04
GEI 2013
SW4-1
3.33
3.50
1.05
GEI 2013
SW4-1
3.33
3.62
1.09
GEI 2013
SW4-1
3.33
4.04
1.22
GEI 2013
SW4-1
3.33
4.08
1.23
GEI 2013
SW4-1
3.33
4.13
1.24
GEI 2013
SW4-1
3.33
4.17
1.25
GEI 2013
SW4-1
3.33
4.34
1.31
GEI 2013
SW4-1
3.33
4.78
1.44
Butler etal. 1993
ST2
3.35
7.00
2.09
GEI 2013
LG1
3.37
3.54
1.05
GEI 2013
LG1
3.37
3.55
1.05
GEI 2013
LG1
3.37
3.90
1.16
GEI 2013
LG1
3.37
3.95
1.17
GEI 2013
LG1
3.37
4.48
1.33
GEI 2013
LG1
3.39
3.00
0.88
GEI 2013
LG1
3.56
2.72
0.77
GEI 2013
LG1
3.56
2.80
0.79
GEI 2013
LG1
3.56
2.89
0.81
GEI 2013
LG1
3.56
2.99
0.84
GEI 2013
LG1
3.56
3.04
0.86
GEI 2013
LG1
3.56
3.08
0.87
GEI 2013
LG1
3.56
3.13
0.88
GEI 2013
LG1
3.56
3.18
0.89
GEI 2013
LG1
3.56
3.25
0.91
GEI 2013
LG1
3.56
3.27
0.92
B-155

-------
White sucker (Catostomus commersonii)
Study

Site
r
^invert
Cfish
Ratio
Butler et al.
1993
WSB2
3.60
4.30
1.19
Butler et al.
1993
WSB2
3.60
6.30
1.75
GEI2013

SWA1
3.64
2.83
0.78
GEI2013

SWA1
3.64
3.39
0.93
GEI 2013

SWA1
3.64
3.47
0.95
GEI 2013

SWA1
3.64
3.55
0.98
GEI 2013

SWA1
3.64
3.63
1.00
GEI 2013

SWA1
3.64
3.75
1.03
Butler et al.
1993
SB2
3.65
4.30
1.18
Butler et al.
1993
R2
3.70
4.20
1.14
Butler et al.
1993
SB2
3.75
4.80
1.28
GEI 2013

CC1
3.76
5.99
1.59
GEI 2013

CC1
3.76
6.56
1.74
GEI 2013

CC1
3.76
7.21
1.92
GEI 2013

CC1
3.76
7.42
1.97
GEI 2013

CC1
3.76
7.62
2.03
Peterson et al. 1991
7
3.83
3.30
0.86
Peterson et al. 1991
7
3.83
4.64
1.21
Butler et al.
1993
R2
3.90
5.40
1.38
Butler et al.
1991
4
3.90
5.30
1.36
GEI 2013

SW88
3.96
4.63
1.17
GEI 2013

SW88
3.96
4.75
1.20
Butler et al.
1993
R1
4.00
9.50
2.38
Butler et al.
1993
ST2
4.10
8.30
2.02
GEI 2013

SW9
4.45
4.07
0.91
GEI 2013

SW9
4.45
4.18
0.94
GEI 2013

SW9
4.45
4.19
0.94
GEI 2013

SW9
4.45
4.20
0.94
GEI 2013

SW9
4.45
4.40
0.99
GEI 2013

SW9
4.45
5.18
1.16
GEI 2013

CC1
4.69
4.51
0.96
GEI 2013

CC1
4.69
4.57
0.98
GEI 2013

CC1
4.69
4.94
1.05
GEI 2013

CC1
4.69
5.02
1.07
GEI 2013

CC1
4.69
5.81
1.24
GEI 2013

CC1
4.69
6.01
1.28
GEI 2013

CC1
4.69
6.43
1.37
GEI 2013

CC1
4.69
7.25
1.55
GEI 2013

CC1
4.69
8.00
1.71
GEI 2013

CC1
4.69
8.52
1.82
Butler et al.
1993
F2
4.80
5.20
1.08
GEI 2013

CC1
5.86
5.00
0.85
B-156

-------
White sucker (Catostomus commersonii)
Study
Site
c
^invert
Cfish
Ratio
GEI2013
CC1
5.86
5.37
0.92
GEI2013
CC1
5.86
5.59
0.95
GEI 2013
CC1
5.86
5.71
0.98
GEI 2013
CC1
5.86
5.90
1.01
GEI 2013
CC1
5.86
6.61
1.13
GEI 2013
CC1
5.86
6.79
1.16
GEI 2013
CC1
5.86
6.82
1.16
GEI 2013
CC1
5.86
7.29
1.25
GEI 2013
CC1
5.86
7.48
1.28
Butler etal. 1991
3
6.20
1.80
0.29
GEI 2013
SWB
7.06
7.18
1.02
GEI 2013
SWB
7.06
7.36
1.04
GEI 2013
SWB
7.06
7.98
1.13
GEI 2013
SWB
7.06
8.03
1.14
GEI 2013
SWB
7.06
9.65
1.37
GEI 2013
SWB
7.06
12.76
1.81
GEI 2013
SWB
7.06
12.85
1.82
GEI 2013
SWB
7.06
13.16
1.86
GEI 2013
SWB
7.44
8.21
1.10
GEI 2013
SWB
7.44
8.77
1.18
GEI 2013
SWB
7.44
8.85
1.19
GEI 2013
SWB
7.44
9.87
1.33
GEI 2013
SWB
7.44
10.97
1.48
GEI 2013
SWB
7.44
13.59
1.83
GEI 2013
SWB
7.44
15.75
2.12
GEI 2013
SWB
7.44
16.40
2.21
Butler etal. 1994
IW
8.35
9.70
1.16
Mueller etal. 1991
R1
8.70
3.40
0.39
GEI 2013
SW2-1
9.14
16.54
1.81
GEI 2013
SW2-1
9.14
18.14
1.99
GEI 2013
SW2-1
9.14
18.54
2.03
GEI 2013
SW2-1
9.14
19.16
2.10
GEI 2013
SW2-1
9.14
21.29
2.33
Lambing et al. 1994
S34
14.00
25.30
1.81
Lambing et al. 1994
S34
14.00
28.00
2.00
Lambing et al. 1994
S34
14.00
29.00
2.07
Butler etal. 1994
HCC1
21.00
3.00
0.14
Butler etal. 1994
GUN2
28.00
20.00
0.71
Butler etal. 1991
7
29.80
7.90
0.27
B-157

-------
White sucker (Catostomus commersonii)
Study Site
c
^invert
Cfish
Ratio
o
o
30
60
Median ratio:
1.11
R2:
0.38
F:
174.4
df:
284
P:
<0.001
Yellow perch (Perca flavescens)
Study
Site
r
^invert
Cfish
Ratio
Butler etal. 1995
PU
0.61
1.10
1.80
Butler etal. 1995
XT
1.07
1.60
1.50
Butler etal. 1995
XT
1.07
1.70
1.60
Butler etal. 1995
MP
1.60
2.00
1.25
Butler etal. 1995
MP
1.60
2.20
1.38
Butler etal. 1995
MP
1.60
2.70
1.69
Belize et al. 2006
Halfway
1.74
2.72
1.56
Belize et al. 2006
Geneva
2.29
3.30
1.44
Belize et al. 2006
Bethel
2.61
3.09
1.19
Belize et al. 2006
McFarlane
3.79
5.40
1.42
Peterson et al. 1991
7
3.83
7.33
1.91
Belize et al. 2006
Long
4.42
6.28
1.42
Belize et al. 2006
Ramsey
4.97
7.64
1.54
Belize et al. 2006
Windy
6.32
6.06
0.96
Belize et al. 2006
Nelson
6.79
10.68
1.57
GEI2013
SW11
8.41
4.54
0.54
GEI2013
SW11
8.41
5.49
0.65
GEI 2013
SW11
8.41
5.50
0.65
GEI 2013
SW11
8.41
5.58
0.66
GEI 2013
SW11
8.41
5.68
0.68
Lambing et al. 1994
S34
14.00
67.00
4.79
B-158

-------
Yellow perch (Perca flavescens)
Median ratio
B-159

-------
Table B-7. Final vertebrate Trophic Transfer Factor (TTF) values, including estimated values using taxonomic classification.
( (iiniiKiii iiiiino
Scientific 11:11110
Order
l-'iiinih
(CIIIIS
III"
I I I-' source
alligator gar
Atractosteus spatula
Lepistosteiformes
Lepisosteidae
Atractosteus
1.21
All fish
black bullhead
Ameiurus melas
Siluriformes
Ictaluridae
Ameiurus
0.85
Exact match
black crappie
Pomoxis nigromaculatus
Perciformes
Centrarchidae
Pomoxis
2.67
Exact match
black redhorse
Moxostoma duquesnei
Cypriniformes
Catostomidae
Moxostoma
1.01
Family Catostomidae
blacknose dace
Rhinichthys atratulus
Cypriniformes
Cyprinidae
Rhinichthys
0.71
Exact match
blue catfish
Ictalurus furcatus
Siluriformes
Ictaluridae
Ictalurus
0.68
Genus Ictalurus
bluegill
Lepomis macrochirus
Perciformes
Centrarchidae
Lepomis
1.03
Exact match
bluehead sucker
Catostomus discobolus
Cypriniformes
Catostomidae
Catostomus
1.04
Exact match
brassy minnow
Hybognathus hankinsoni
Cypriniformes
Cyprinidae
Hybognathus
1.20
Family Cyprinidae
brook stickleback
Culaea inconstans
Gasterosteiformes
Gasterosteidae
Culaea
1.79
Exact match
brook trout
Salvelinus fontinalis
Salmo niformes
Salmonidae
Salvelinus
0.88
Exact match
brown bullhead
Ameiurus nebulosus
Siluriformes
Ictaluridae
Ameiurus
0.85
Genus Ameiurus
brown trout
Salmo trutta
Salmo niformes
Salmonidae
Salmo
1.38
Exact match
bullhead

Siluriformes
Ictaluridae

0.77
Family Ictaluridae
burbot
Lota lota
lota
Gadiformes
Lotidae
1.21
All fish
chain pickerel
Esox niger
Esociformes
Esocidae
Esox
1.78
Genus Esox
channel catfish
Ictalurus punctatus
Siluriformes
Ictaluridae
Ictalurus
0.68
Exact match
common carp
Cyprinus carpio
Cypriniformes
Cyprinidae
Cyprinus
1.20
Exact match
common snook
Centropomus undecimalis
Perciformes
Centropomidae
Centropomus
1.41
Order Perciformes
crappie
Pomoxis sp.
Perciformes
Centrarchidae
Pomoxis
2.67
Genus Pomoxis
creek chub
Semotilus atromaculatus
Cypriniformes
Cyprinidae
Semotilus
1.06
Exact match
cutthroat trout
Oncorhynchus clarkii
Salmo niformes
Salmonidae
Oncorhynchus
1.12
Exact match
fltvi 0 r
desert pupfish
Cyprinodon macularius
Cyprinodontiformes
Cyprinodontidae
Cyprinodon
1.24
UIUCI
Cyprinodontiformes
dolly varden
Salvelinus malma
Salmo niformes
Salmonidae
Salvelinus
0.88
Genus Salvelinus
fathead minnow
Pimephales promelas
Cypriniformes
Cyprinidae
Pimephales
1.57
Exact match
flannelmouth sucker
Catostomus latipinnis
Cypriniformes
Catostomidae
Catostomus
0.98
Exact match
flathead catfish
Pylodictis olivaris
Siluriformes
Ictaluridae
Pylodictus
0.77
Family Ictaluridae
flathead chub
Platygobio gracilis
Cypriniformes
Cyprinidae
Platygobio
1.20
Family Cyprinidae
freshwater drum
Aplodinotus grunniens
Perciformes
Sciaenidae
Aplodinotus
1.41
Order Perciformes
gizzard shad
Dorosoma cepedianum
Clupeiformes
Clupeidae
Dorosoma
1.21
All fish
goldeye
Hiodon alosoides
Hiodontiformes
Hiodontidae
Hiodon
1.21
All fish
green sunfish
Lepomis cyanellus
Perciformes
Centrarchidae
Lepomis
1.12
Exact match
iowa darter
Etheostoma exile
Perciformes
Percidae
Etheostoma
1.51
Family Percidae
kokanee salmon
Oncorhynchus nerka
Salmo niformes
Salmonidae
Oncorhynchus
1.10
Genus Oncorhynchus
largemouth bass
Micropterus salmoides
Perciformes
Centrarchidae
Micropterus
1.39
Exact match
largescale sucker
Catostomus macrocheilus
Cypriniformes
Catostomidae
Catostomus
1.01
Genus Catostomus
longnose dace
Rhinichthys cataractae
Cypriniformes
Cyprinidae
Rhinichthys
0.71
Genus Rhinichthys
B-160

-------
( (iiniiKiii iiiiino
Scientific 11:11110
Order
l;uilil\
(ciiiis
III"
I I I-' source diilii
longnose sucker
Catostomus catostomus
Cypriniformes
Catostomidae
Catostomus
0.90
Exact match
mottled sculpin
Cottus bairdi
Scorpaeniformes
Cottidae
Cottus
1.38
Exact match
mountain whitefish
Prosopium williamsoni
Salmoniformes
Salmonidae
Prosopium
1.38
Exact match
nine spine stickleback
Pungitius pungitius
pungitius
Gasterosteiformes
Gasterosteidae
1.79
Family Gasterosteidae
northern pike
Esox lucius
Esociformes
Esocidae
Esox
1.78
Exact match
northern pikeminnow
Ptychocheilus oregonensis
Cypriniformes
Cyprinidae
Ptychocheilus
1.20
Family Cyprinidae
northern plains killifish
Fundulus kansae
Cyprinodontiformes
Fundulidae
Fundulus
1.27
Exact match
northern redbelly dace
Chrosomus eos
Cypriniformes
Cyprinidae
Chrosomus
1.20
Family Cyprinidae
northern squawfish
Ptychocheilus oregonensis
Cypriniformes
Cyprinidae
Ptychocheilus
1.20
Family Cyprinidae
quillback
Carpiodes cyprinus
Cypriniformes
Catostomidae
Carpiodes
1.01
Family Catostomidae
rainbow trout
Oncorhynchus mykiss
Salmoniformes
Salmonidae
Oncorhynchus
1.07
Exact match
razorback sucker
Xyrauchen texanus
Cypriniformes
Catostomidae
Xyrauchen
1.01
Family Catostomidae
red shiner
Cyprinella lutrensis
Cypriniformes
Cyprinidae
Cyprinella
1.31
Family Cyprinidae
redbreast sunfish
Lepomis auritus
Perciformes
Centrarchidae
Lepomis
1.07
Genus Lepomis
redear sunfish
Lepomis microlophus
Perciformes
Centrarchidae
Lepomis
1.07
Genus Lepomis
redside shiner
Richardsonius balteatus
Cypriniformes
Cyprinidae
Richardsonius
1.08
Exact match
river carpsucker
Carpiodes carpio
Cypriniformes
Catostomidae
Carpiodes
1.01
Family Catostomidae
river redhorse
Moxostoma carinatum
Cypriniformes
Catostomidae
Moxostoma
1.01
Family Catostomidae
rock bass
Ambloplites rupestris
Perciformes
Centrarchidae
Ambloplites
1.12
Family Centrarchidae
roundtail chub
Gila robusta
Cypriniformes
Cyprinidae
Gila
1.20
Family Cyprinidae
sacramento perch
Archoplites interruptus
Perciformes
Centrarchidae
Archoplites
1.12
Family Centrarchidae
sacramento pikeminnow
Ptychocheilus grandis
Cypriniformes
Cyprinidae
Ptychocheilus
1.20
Family Cyprinidae
sailfin molly
Poecilia latipinna
Cyprinodontiformes
Poeciliidae
Poecilia
1.21
Family Poeciliidae
sand shiner
Notropis stramineus
Cypriniformes
Cyprinidae
Notropis
1.56
Exact match
sauger
Sander canadensis
Perciformes
Percidae
Sander
1.60
Genus Sander
sculpin
Cottus sp.
Scorpaeniformes
Cottidae
Cottus
1.29
Exact match
shadow bass
Ambloplites ariommus
Perciformes
Centrarchidae
Ambloplites
1.12
Family Centrarchidae
shorthead redhorse
Moxostoma macrolepidotum
Cypriniformes
Catostomidae
Moxostoma
1.01
Family Catostomidae
silver carp
Hypophthalmichthys molitrix
Cypriniformes
Cyprinidae
Hypophthalmichthys
1.20
Family Cyprinidae
smallmouth bass
Micropterus dolomieu
Perciformes
Centrarchidae
Micropterus
0.86
Exact match
smallmouth buffalo
Ictiobus bubalus
Cypriniformes
Catostomidae
Ictiobus
1.01
Family Catostomidae
speckled dace
Rhinichthys osculus
Cypriniformes
Cyprinidae
Rhinichthys
0.71
Genus Rhinichthys
spottail shiner
Notropis hudsonius
hudsonius
Cypriniformes
Cyprinidae
1.56
Genus Notropis
spotted bass
Micropterus punctulatus
Perciformes
Centrarchidae
Micropterus
1.12
Genus Micropterus
spotted gar
Lepisosteus oculatus
Lepistosteiformes
Lepisosteidae
Lepisosteus
1.21
All fish
stonecat
Noturus flavus
Siluriformes
Ictaluridae
Noturus
0.77
Family Ictaluridae
striped bass
Morone saxatilis
Perciformes
Moronidae
Morone
1.48
Exact match
striped mullet
Mugil cephalus
Mugiliformes
Mugilidae
Mugil
1.21
All fish
B-161

-------
( (iiniiKiii iiiiino
Scientific 11:11110
Order
l-'iiinih
(ciiiis
III"
I I I-' source diilii
sucker

Cypriniformes
Catostomidae

1.01
Family Catostomidae
tilapia

Perciformes
Cichlidae

1.41
Order Perciformes
trout species
Oncorhynchus sp.
Salmoniformes
Salmonidae
Oncorhynchus
1.10
Genus Oncorhynchus
tui chub
Gila bicolor
Cypriniformes
Cyprinidae
Gila
1.20
Family Cyprinidae
utah sucker
Catostomus ardens
Cypriniformes
Catostomidae
Catostomus
1.01
Genus Catostomus
walleye
Sander vitreus
Perciformes
Percidae
Sander
1.60
Exact match
western mosquitofish
Gambusia affinis
Cyprinodontiformes
Poeciliidae
Gambusia
1.21
Exact match
white bass
Morone chrysops
Perciformes
Moronidae
Morone
1.48
Genus Morone
white crappie
Pomoxis annularis
Perciformes
Centrarchidae
Pomoxis
2.67
Genus Pomoxis
white sturgeon
Acipenser transmontanus
Acipenseriformes
Acipenseridae
Acipenser
1.21
All fish
white sucker
Catostomus commersonii
Cypriniformes
Catostomidae
Catostomus
1.11
Exact match
wiper
Morone chrysops x Moron saxatilis
Perciformes
Moronidae
Morone
1.48
Genus Morone
yellow perch
Perca Jlavescens
Perciformes
Percidae
Perca
1.42
Exact match
B-162

-------
4.0 Food Web Models Used to Calculate Composite TTFs to Translate the Egg-Ovary FCV to Water-Column Values
Table B-8. Food web models used to calculate composite TTFs to translate the egg-ovary FCV to a water-column value at aquatic sites where sufficient data was available to calculate an enrichment factor (EF).
Ki-li-mi
Sili-
Sili-
T;iri-l
I'ish
I'ish pri-\ :is di-si rihi-d in \;iiuri-Si-r\i- I'ish piv\ spp
r H2 in
1 11.2
1
r 11.2
2 11.2
">o
2
2 11.2
3 11.2
3
3
3 11.2 4 11.2 4 4 4 11.2
1
r 11 .3
r Ti.3
r r ti.3
I! Hill i\
T;iri-
TTI''i'iuii|>iisi
ii-
ik'siripi
ll)
11 sli
III
iiuii nii-nl
spi-iii-s usi-d
S|)|)
11.2
l>i-i-n|>i>rl in
S|)|) llsi-d
11.2
11.2
priipniiiii s|>|> usi-d 11.2 TI.2 prupiniiii
11.3
S|)|) lisi-il
S|)|)
11.3 pnipiirliii
I- I I I "
1 llsli
li-

inn

spi-iii-s



;il>l>ri-\
II
n

S|)|)
II
n

S|)|)
II
11 S|)|) II 11
S|)|)

;il>l>ri-\
II n

III




i n in mi i




1


;il>ri-\
1


;il>ri-\
1
;il>ri-\ I-"



1






n n;inii-





















Default


black
0.85
Omnivorous bottom feeder; often eats
Median of all
in
2.14
0.45
Median of
CIS
1.41
0.45
Median of
bvs
4.29
0.10




2.03
0.85
1.72



bullhea

aquatic insects, crustaceans, molluscs,
insects



all



all













d

occasionally fishes and carrion




crustacean



bivalves










Default


black
2.67
Primarily a midwater feeder; zooplankton
Median of all
in
2.14
0.50
S
Median of
pc
1.41
0.10




Fish
Median all
f+a
2.28 0.4
2.12
2.67
5.66



crappie

and small Diptera larvae predominate in
insects



planktonic








fish eating










the diet of individuals to 12 cm SL, while




crustacean








median all










fishes and aquatic insects predominate in




s








invertebrat










the diet of larger individuals













es





Default


blackno
0.71
Eats immature aquatic insects,
Median of all
in
2.14
0.50
Median of
all
1.48
0.50








1.81
0.71
1.29
Default
Default
Default
Default
bluegill 1.03
se dace	amphipods, and various other aquatic
invertebrates; also eats algae and
diatoms, which may be of little
nutritional value (Smith 1979, Becker
1983).
blue	0.68 Bottom feeder. Eats mostly crustaceans
catfish	and aquatic insects when young. Later,
fishes and large invertebrates become
most important (Moyle 1976). Also
scavenges.
Feeds opportunistically on aquatic insect
larvae, planktonic crustaceans, flying
insects, snails, and other small
invertebrates; small fishes, fish eggs,
crayfish, and algae sometimes are eaten.
Larvae and juveniles often eat
cladocerans and copepod nauplii. Adults
eats mainly aquatic insects, crayfishes,
and small fishes, or, in some bodies of
water, mostly zooplankton. Feeds at all
levels of water column,
bluehea 1.04 Herbivore, Invertivore
d
sucker
brassy 1.20 Eats algae, phyto- and zooplankton,
minnow	benthic invertebrates, surface drift,
bottom ooze (Becker 1983).
insects
Median of all in
insects
Median of all in
insects
all
invertebrat
es except
bivalves
2.14
2.14
0.36 Median
insects and
benthic
crustacean
s
0.68 Median of pc
planktonic
crustacean
s
in,be 1.74
1.41
0.20 Median of bvs
all
bivalves
0.20 crayfish cr
4.29
1.46
0.08
0.08
TL1
TL1
TL1
TL1
1.00
1.00
0.60 Median of all
all
invertebrat
es except
bivalves
0.50 Median of pc
planktonic
crustacean
1.48
1.41
0.40
0.40 Median of in
all insects
2.14
0.10
Fish Median all f+a 2.28	0.36
fish eating
median all
invertebrat
es
Fish Median all f+a 2.28	0.04
fish eating
median all
invertebrat
es
2.28 0.6
1.56
1.95 1.03	2.00
1.19 1.04	1.24
1.28 1.20	1.54
B-163

-------
Ki-l'i-mi Sili-	Sili-
ii-	di-siripi II)
inn
T;iri-l I'isli
iisii	in
spi-iii-s
imiiinii
11 11:11111-
I'isli |)n\ :is di-si rihi-d in \:iliiri-Si-r\i-
i"ish pn-\ spp 1 11.2111 i ti.2 r	rn.2 2 11.2
I'lininii-nl	spi-iii-s usi-d spp	11.2	pnipiiriin spp usi-d
;il>l>ri-\ II	n
I
">o
11.2
spp
;il)iv\
Default
Default
Default
Default
Default
Default
Default
brook	1.79 Eats various aquatic invertebrates
stickleb	(including eggs and larvae), eggs and
ack	larvae of fishes, and algae. In a Manitoba
lake, was opportunistic but heavily
dependent on arthropods (Moodie 1986).
brook 0.88 Feeds opportunistically on various
trout	invertebrate and vertebrate animals,
including primarily terrestrial and aquatic
insects and planktonic crustaceans.
brown 0.85 Bottom feeder. Young eat chironomid
bullhea	larvae and small crustaceans. Adults eat
d	larger insect larvae and fishes, also fish
eggs, mollusks, carrion, and plant
material (Becker 1983, Moyle 1976).
brown 1.38 Eats aquatic and terrestrial insects and
trout	their larvae, crustaceans (especially
crayfish), molluscs, fishes, and other
animals. In streams, young feed mainly
on aquatic and terrestrial drift
invertebrates; in lakes, they feed on
zooplankton and benthic invertebrates
(Sublette et al. 1990). Large adults feed
on fishes, crayfish, and other benthic
invertebrates.
Black (not exotic to CO and NM):
Omnivorous bottom feeder; often eats
aquatic insects, crustaceans, molluscs,
occasionally fishes and carrion. Stomach
often contain substantial amounts of plant
material of unknown nutritional value
(Moyle 1976). Juveniles planktivorous; at
about 27 mm TL, feed largely on
crustaceans and midge larvae
1.21 Young eat mainly immature aquatic
insects, crayfish, molluscs, and other
deepwater invertebrates. Larger
individuals feed mostly on fishes (Becker
1983, Scott and Crossman 1973).
0.68 Bottom feeder. Young eat mainly small
invertebrates; as they grow, fishes and
crayfish become increasingly important,
though individuals of all sizes eat
abundant aquatic insects. Large fish are
mainly piscivorous (Moyle 1976).
bullhea 0.77
d
burbot
channel
catfish
Median of all all
invertebrates
except
bivalves
Median of all in
insects
Median of all in
insects
Median of
planktonic
crustaceans
pc
1.48
2.14
2.14
1.41
0.80 TL1
TL1
0.60 crayfish cr
0.68 Median of
all
invertebrat
es except
bivalves
0.20 Median of
all insects
Median of all in
insects
2.14
Median of all in
insects
Median of all in
insects
2.14
2.14
all
0.68 Median of
all
invertebrat
es except
bivalves
0.25 Median of
all
crustacean
s
0.48 Median of
planktonic
crustacean
all
pc
2 2 11.2 3 11.2 3 3 3 11.2 4 11.2 4 4 4 11.2 1 1 11.3
TI.2 pnipiiriin spp usi-d I I.2 TI.2 pnipiiriin spp usi-d I I.2 TI.2 pnipiirliii TI.3 spp usi-d
n	spp TT n	spp TT n	spp
II
I
:il>ri-\ !"
II
;il>ri-\ T
1 TI.3 1	l TI.3
spp T1.3	pniporiio
;il>l>ri-\ 'IT	n
I
IH'li-iiiN T:iri- I Tl'iiiniposi
i- I TI-" 1 fish k-
III
1.00
0.20
1.38 1.79
2.47
1.46	0.10 Median of bvs 4.29	0.05
all
bivalves
1.48	0.20 Median of bvs
all
bivalves
4.29
0.04
Fish Median all f+a 2.28
fish eating
median all
invertebrat
es
Fish Median all f+a 2.28
fish eating
median all
invertebrat
0.25	2.22 0.S
1.96
0.08	2.11 0.85	1.79
2.14
0.12 crayfish cr
1.46
0.08
Fish Median all f+a
fish eating
median all
invertebrat
2.28
0.6	2.02 1.38	2.78
1.48	0.20 Median of bvs 4.29	0.04
all
bivalves
Fish Median all
fish eating
median all
invertebrat
f+a
2.28	0.08
2.11 0.77	1.62
1.41	0.25
1.41	0.20 crayfish cr	1.46	0.08
Fish Median all f+a
fish eating
median all
invertebrat
es
Fish Median all f+a
fish eating
median all
invertebrat
es
2.28 0.5
2.28	0.24
2.03 1.21	2.45
1.97 0.6
1.35
B-164

-------
Ki-l'i-ri-n Sili-	Sili-
ii-	di-siripi II)
inn
T;iri-l I'isli
iisii	in
spi-iii-s
imiiinii
11 11:11111-
I'isli |)n\ :is di-si rihi-d in \:iluri-Si-r\i-
i"ish pn-\ spp 1 11.2111 i ti.2 r rn.2 2 11.2
imiimi-nl	spi-iii-s usi-d spp	11.2 prupurliii spp usi-d
;il>l>ri-\ II n
">o
11.2
spp
;il>ri-\
I	1.2
II
I
2 I 1.2
pnipiiriin
n
3 11.2
spp iim-(I
I 1.2
spp
:il>ri-\
11.2
II
I
3 11.2
pnipiiriin
n
4 11.2
spp IISI-(I
4
11.2
spp
;il>ri-\
4
n.2
ii
i
4 11.2
pnipiiriin
n
1
11.3
spp
r 11 .3
spp IISI-(I
rn.3 r
spp 11.3
;il>l>ri-\ IT
I
1  TI.3	IH'li-iiiN T:iri- Tl Tiiimposi
pnipiiriin	i- III I llsli k-
n	III
Default
Default
Default
Default
Default
Default
Default
Default
commo
n carp
crappie
creek
chub
cutthroa
t trout
fathead
minnow
flannel
mouth
sucker
flathead
chub
freshwa
ter
drum
1.20
2.67
1.06
1.12
1.57
0.98
1.20
1.41
Omnivorous; adults eat mainly
invertebrates, detritus, fish eggs, and
plant material (Jester 1974, Becker 1983,
Sublette et al. 1990).
Black: Primarily a midwater feeder;
zooplankton and small Diptera larvae
predominate in the diet of individuals to
12 cm SL, while fishes and aquatic
insects predominate in the diet of larger
individuals. White: eats fishes, planktonic
crustaceans, and aquatic insects; small
individuals eat mostly zooplankton, fish
tend to predominate in the diet of larger
individuals, though zooplankton also
consumed (Moyle 1976)
Feeds opportunistically on various plants
and animals, from surface drift to
benthos; mostly invertivorous but large
individuals often picivorous (Becker
1983). Chironomid larvae and other
larval insects important in diet of young.
Opportunistic. Inland cutthroats feed
primarily on insects (aquatic and
terrestrial); often feeds in and especially
downstream from riffle areas; some large
individuals feed mostly on fishes; also
eats zooplankton and crustaceans.
Feeds opportunistically in soft bottom
mud; eats algae and other plants, insects,
small crustaceans, and other invertebrates
(Becker 1983, Sublette et al. 1990).
Herbivore, Invertivore Bottom feeder.
Reported to feed on diatoms, algae,
fragments of higher plants, seeds, and
benthic invertebrates (Sigler and Miller
1963; Lee et al. 1980). See Tyus and
Minckley 1988 for possible importance
of Mormon cricket as food source.
Opportunistic; eats aquatic and terrestrial
insects and some algae (Olund and Cross
1961)
Young feed mainly on minute
crustaceans; adults mostly are bottom
feeders, eat insect larvae, crustaceans,
fishes, and (mostly in rivers) clams and
expected diet
of small
invertebrates
Median of all
invertebrates
except
bivalves
Median of all
insects
all
1.48
2.14
0.65 TL1
Median of all all
invertebrates
except
bivalves
Median of all in
insects
Median of all in
insects
Median	in,be
insects and
benthic
crustaceans
Median of all in
insects
Median of all crs
crustaceans
1.48
2.14
2.14
1.74
2.14
1.41
TL1 1.00
0.50 Median of pc
planktonic
crustacean
1.41
0.35
0.10
Fish Median all
fish eating
median all
invertebrat
0.70 TL1
TL1 1.00
0.20
0.50 Median of crs
all
crustacean
s
0.60 Median of crs
all
crustacean
1.41
0.20
0.75 TL1
TL1
1.41
1.00
0.20 TL1
0.25
TL1 1.00
0.20
0.80 TL1
TL1 1.00
0.44 Median of in 2.14
all insects
0.20
0.40 Median of bvs 4.29
all
bivalves
0.04
f+a
2.28
0.4
1.31 1.20	1.58
2.12 2.67	5.66
Fish Median all
fish eating
median all
invertebrat
es
Fish Median all
fish eating
median all
invertebrat
es
f+a
2.28
0.1
1.46
f+a
2.28
0.3
2.04
1.77
1.55
Fish Median all f+a
fish eating
median all
invertebrat
2.28
0.12
1.06
1.12
1.57
0.98
1.55
2.29
2.78
1.52
1.91 1.20	2.30
1.92 1.41	2.71
B-165

-------
Ki-li-ri-n Sili-	Sili- T;iri-l	I'isli I'isli pri-\ :is di-snihi-d in \:iliiri-Si-r\i- I'isli pri-> spp 1 TI.2 TTI-" lTI.2	1	lTI.2	2 TI.2 2
ii-	di-siripi II)	IIsli	nr	iiiiiiiiii-nl	spi-iii-s usi-d spp	11.2	pnipuriiii	spp usi-d 11.2
inn	spi-iii-s	;il>l>ri-\	II	n	spp
I'lininiii	I'	;il>ri-\
n 11:11111-
snails (Becker 1983, Scott and Crossman
1973, Lee etal. 1980).
Default
Default
Default
Default
Default
Default
Default
Default
gizzard 1.21 Adults primarily bottom filter-feeding
shad	detritivores
goldeye 1.21 Young-of-year eat mainly
microcrustaceans, also other
invertebrates. Older individuals eat
mainly aquatic insects obtained at surface
but also various other animals, including
frogs, fishes, and small mammals,
green	1.12 Feeds opportunistically on the larger,
sunfish	more active invertebrates that occur with
them, and on small fishes. Young feed
mostly on crustaceans (zooplankton) and
aquatic insect larvae. Adults eat more
large aquatic and terrestrial insects,
crayfish, and fishes
Iowa	1.51 Eats mainly various invertebrates;
darter	commonly ingested food items of adults
are midge larvae, mayfly naiads, and
amphipods, and of the young, copepods
and cladocerans. Apparently feeds on
swimming organisms and those on
bottom.
1.10 Zooplankton, insects.
kokane
e
salmon
largemo
uth bass
expected diet
of small
invertebrates
1.39 Fry feed mainly on zooplankton. Larger
young eat insects, crustaceans, and fish
fry. Adults eat mainly fishes, though
sometimes prefer crayfish or amphibians
(Moyle 1976, Smith 1979).
longnos 0.71 Eats mainly benthic insects, especially
e dace	Diptera and mayflies (Becker 1983, Scott
and Crossman 1973); also eats algae and
plant material (Sublette et al. 1990).
Terrestrial insects and fish eggs common
in diet of adults from Lake Michigan (see
Sublette et al. 1990).
longnos 0.90 Eats mostly bottom invertebrates (Scott
e sucker	and Crossman 1973).
TL1	TL1
Median	in,be
insects and
benthic
crustaceans
Median of all in
insects
1.00
1.74
1.00
1.00
2.14
Median of all in
insects
2.14
0.58 Median of pc
planktonic
crustacean
0.70 amphipods am
Median of
planktonic
crustaceans
pc
Median of all in
insects
Median of all in
insects
1.41
2.14
2.14
0.80 Median of in
all insects
0.10 crayfish cr
0.80 TL1
TL1
Median of all
invertebrates
except
bivalves
all
1.48
1.00
11.2
II
I
2 11.2
pnipuriiii
11
3 11.2
spp usi-il
I 1.2
spp
:il>ri-\
11.2
II
I
3 11.2
pnipuriiii
11
4 11.2
spp IISI-(I
4
11.2
spp
;il>ri-\
4
11.2
IT
I
4 11.2
pnipuriiii
11
r rn.3
I 1.3 spp llsi-(l
spp
rn.3 r	in.3
spp	11.3	pnipuriiii
;il>l>ri-\ 'IT	11
I
IH'li-iiiN T:iri- I Tl'iiinipnsi
i- I TI-" 1 fish k-
III
1.00
1.74
1.21
1.21
1.21
2.10
1.41
0.10 crayfish cr
1.46
0.08
Fish Median all
fish eating
median all
invertebrat
es
f+a
2.28
0.24
2.05
1.12
2.29
1.22	0.16 crayfish cr	1.46	0.08 Median of pc 1.41	0.06	1.90 1.51	2.87
planktoni
c
Crustacea
ns
2.14
0.20
1.56
1.10
1.71
1.46	0.10	Fish Median all f+a 2.28	0.8	2.18 1.39	3.04
fish eating
median all
invertebrat
es
1.00	0.20	1.91 0.71	1.36
1.48 0.90	1.34
B-166

-------
Ki-l'i-ri-n Sili-	Sili-
ii-	di-siripi II)
inn
T;iri-l I'isli
iisii	in
spi-iii-s
imiiinii
11 11:11111-
I'isli |)n\ :is di-si rihi-d in \:iluri-Si-r\i-
i"ish pn-\ spp 1 11.2111 i ti.2 r rn.2 2 11.2
I'lininii-nl	spi-iii-s usi-d spp	11.2 prnpnrlin spp usi-d
;il>l>ri-\ II n
">o
11.2
spp
;il>ri-\
I	1.2
II
I
2 I 1.2
prnpnrlin
n
3 11.2
spp iim-(I
I 1.2
spp
:il>ri-\
3	3 11.2	4 11.2 4	4	4 11.2	1 1 11.3
TI.2	pnipiiriin spp usi-d I I.2	TI.2	pnipiiriin	TI.3 spp usi-d
II	n	spp	TT	n	spp
T	;il)ri-\	I-"
rn.3 r	in.3
spp TI.3	pnipiiriin
;il>l>ri-\ 'IT	n
I
IH'li-iiiN T:iri- I Tl'i'iinipiisi
i- I TI-" 1 llsli k-
III
Default
Default
Default
Default
Default
mottled 1.38 lieiitliic feeder; forages among rocks,
sculpin	mainly on immature aquatic insect larvae,
especially mayflies, chironomid midges,
and stoneflies; larger individuals also eat
caddisflies and crayfish; crustaceans,
annelids, fishes (including fish eggs) and
plant material also may be eaten; may
take swimming prey from water column
(Scott and Crossman 1973, Becker 1983).
mountai 1.38 Feeds actively on aquatic and terrestrial
n	insects. Also feeds on some fish eggs and
whitefis	occasionally on fishes. Bottom-oriented
h	predator (Moyle 1976), occasionally
feeds at surface (Sigler and Sigler 1987).
ninespi 1.79 Eats mainly small crustaceans and
ne	aquatic insects; sometimes also fish eggs
stickleb	and fry (Becker 1983).
ack
norther 1.78 Young initially eat large zooplankton and
n pike	immature aquatic insects. After 7-10 days
fishes begin to enter diet and eventually
dominate. Adults feed opportunistically
on vertebrates small enough to be
engulfed. (Scott and Crossman 1973).
Sight feeder.
norther 1.27 Feed effectively at all levels and food
n plains	habits are generalized. Prefer aquatic
killifish	insects but also feed on plants.
Median of all in	2.14
insects
Median of all in
insects
Median of all crs
crustaceans
Montana field
guide
(http://fieldgui
de.mt.gov/detai
1_AFCNB0460
O.aspx)
Median
insects and
benthic
crustaceans
Median of all
insects
in,be
2.14
1.41
1.74
2.14
0.70
Median of
all
crustacean
s
0.90
0.48 Median of
all insects
0.05
0.80 TL1
TL1
1.41
2.14
1.00
0.10 TL1
0.44
0.20
TLl
1.00
0.10
l'ish Median all f+a
fish eating
median all
invertebrat
Fish Median all f+a
fish eating
median all
invertebrat
es
Fish Median all f+a
fish eating
median all
invertebrat
es
Fish Median all f+a
fish eating
median all
invertebrat
es
2.28
2.28
2.28
2.28
0.1
0.1
0.08
0.95
1.97 1.38	2.72
2.16
2.25
1.91
1.38
1.79
1.78
1.27
2.97
3.22
4.02
2.44
Default
norther
n
redbelly
dace
1.20 Eats mainly diatoms and filamentous
algae, also zooplankton and aquatic
insects.
TLl
TLl
1.00
0.70
Median of
all insects
2.14
0.15 Median
insects
and
benthic
crustacean
s
in,be
1.74
0.15
1.28
1.20
1.54
Default
norther 1.20 Small individuals feed primarily on
n	aquatic and terrestrial insects. Adults feed
squawfi	on fish, insects, insect larvae, crustaceans
sh	and some plankton during spring and
summer. Fishes are the major component
Median of all in
insects
2.14
0.32
Median of
all
crustacean
s
1.41
B-167
0.08
Fish Median all
fish eating
median all
invertebrat
f+a
2.28
0.6
2.17
1.20
2.61

-------
Ki-li-ri-n Sili-	Sili- T;iri-l	I'isli I'ish pri-\ :is di-sirihi-d in \:iliiri-Si-r\i- I'isli pri-> spp 1 TI.2 TTI-" lTI.2	1	lTI.2	2 TI.2 2
ii-	di-siripi II)	IIsli	nr	iiiiiiiiii-nl	spi-iii-s usi-d spp	11.2	pnipuriiii	spp usi-d 11.2
inn	spi-iii-s	;il>l>ri-\	II	n	spp
I'lininiii	I'	;d>ri-\
n 11:11111-
of the diet in winter.
Default
Default
Default
rainbow 1.07 In lakes, feeds mostly on bottom-
trout	dwelling invertebrates (e.g., aquatic
insects, amphipods, worms, fish eggs,
sometimes small fish) and plankton. In
streams, feeds primarily on drift
organisms. May ingest aquatic vegetation
(probably for attached invertebrates),
red	1.31 Eats various small invertebrates (insects,
shiner	crustaceans), plant material (digestibility
may be low), and microorganisms
(Becker 1983). In Virgin River, diet
dominated by Ceratopongidae,
Simuliidae, and Chironomidae (Greger
and Deacon 1988).
redside 1.08 Feeds mainly on aquatic and terrestrial
shiner	insects; also eats molluscs, plankton, and
some small fish and fish eggs. Fry eat
zooplankton and algae.
Median of all in	2.14
insects
0.75
Median	in,be 1.74
insects and
benthic
crustaceans
Median of all in	2.14
insects
1.00
0.70 Median of
planktonic
crustacean
pc
Default
Default
river	1.01 Mo stly a bottom feeder, browses on
carpsuc	periphyton associated with submerged
ker	rocks and debris, ingests various small
planktonic plants and animals,
roundta 1.20 Opportunistic; eats available aquatic and
il chub	terrestrial insects, gastropods,
crustaceans, fishes, and sometimes
filamentous algae (Sublette et al. 1990).
TL1
TL1 1.00
Median of all in
insects
2.14
0.75
0.55
Median of
planktonic
crustacean
s
Median of
all
crustacean
s
pc
Default
Default
Sacram 1.12 Opportunistic; diet mainly benthic insect
ento	larvae, snails, mid-water insects,
perch	zooplankton, and fishes (Moyle et al.
1989). Young feed mainly on small
crustaceans, but as they grow Sacramento
perch consume more aquatic insect larvae
and pupae. Large adults feed mainly on
other fishes when available,
sailfin 1.21 Eats mainly algae, vascular plants,
molly	organic detritus, and mosquito larvae
(and other small invertebrates).
TL1
TL1
1.00
TL1
TL1 1.00
0.75 Median of
all insects
0.75 Median of
all insects
11.2
II
I
2 11.2
pnipuriiii
11
3 11.2
spp usi-d
I 1.2
spp
:d>ri-\
11.2
II
I
3 11.2
pnipuriiii
11
4 11.2
spp usi-d
4
11.2
spp
;d>ri-\
4
11.2
IT
I
4 11.2
pnipuriiii
11
1
11.3
spp
r 11.3
spp usi-d
1 TI.3 1
spp T1.3
;il>l>ri-\ IT
I
1 I 1.3
pnipuriiii
11
I'. I'li-ii i\
i- I I I "
T:iri-
I llsli
III
iri'i'iinipnsi
II-
Fish Median all f+a 2.28	0.25	2.18 1.07	2.33
fish eating
median all
invertebrat
es
1.74 1.31	2.27
1.41
1.41
0.10
0.25
Median of
all
bivalves
bvs
4.29
0.10
Fish Median all
fish eating
median all
invertebrat
es
f+a
2.28
0.1
2.30
1.10
1.08
1.01
2.48
1.11
1.41	0.15 Median of bvs 4.29	0.15
all
bivalves
2.14	0.25
Fish Median all f+a 2.28	0.15	2.38 1.20	2.86
fish eating
median all
invertebrat
es
1.29 1.12	1.44
2.14	0.25
1.29 1.21	1.56
B-168

-------
Ki-l'i-ri-n Sili-	Sili-
ii-	di-siripi II)
inn
T;iri-l I'isli
iisii	in
spi-iii-s
imiiinii
11 11:11111-
I'isli |)n\ :is di-si rihi-d in \:iluri-Si-r\i-
i"ish pn-\ spp 1 11.2111 i ti.2 r	rn.2 2 11.2
I'lininii-nl	spi-iii-s usi-d spp	11.2	prnpnrlin spp usi-d
;il>l>ri-\ II	n
I
">o
11.2
spp
;il>ri-\
Default
Default
Default
Default
Default
sand	1.56 Eats various aquatic and terrestrial
shiner	invertebrates (especially chironomids),
algae, and (mainly) bottom particulate
matter (Becker 1983). Winter diet mostly
chironomids larvae and mayfly and
stonefly naiads (Ohio, see Sublette et al.
1990)
sauger 1.60 Larvae eat microcrustaceans. Young eat
zooplankton, immature and adult aquatic
insects, and fish fry; adults eat small
fishes and various invertebrates (Scott
and Crossman 1973), or are almost
exclusively piscivorous (Burkhead and
Jenkins 1991). Sight feeder, adapted to
low light.
sculpin 1.29 Benthic feeder; forages among rocks,
mainly on immature aquatic insect larvae,
especially mayflies, chironomid midges,
and stoneflies; larger individuals also eat
caddisflies and crayfish; crustaceans,
annelids, fishes (including fish eggs) and
plant material also may be eaten; may
take swimming prey from water column
(Scott and Crossman 1973, Becker 1983).
1.01 Invertivore
shorthe
ad
redhors
e
smallm
outh
bass
0.86 Adults almost entirely piscivorous if
sufficient prey available
Median	in,be
insects and
benthic
crustaceans
Median	in,be
insects and
benthic
crustaceans
Median of all in
insects
1.74
0.75 TL1
1.74
2.14
Median of all all
invertebrates
except
bivalves
Median of all in
insects
1.48
2.14
TL1
0.36 Median of pc
planktonic
crustacean
0.70 crayfish
1.00
0.20
Default
speckle 0.71 An omnivorous benthic feeder, at times
d dace	feeding on drift in mid-water or rarely at
the surface (Schreiber and Minckley
1981). The diet consists mostly of benthic
insects, also includes other invertebrates,
algae, and detritus (little or no plant
material or detritus in some areas)
(Sublette et al. 1990, Woodbury 1933,
Greger and Deacon 1988). Young feed
mainly on zooplankton.
Median of all in	2.14
insects
0.70 Median in,be
insects and
benthic
crustacean
s
11.2
II
I
2 11.2
prnpnrlin
n
3 11.2 3
spp usi-d I I.2
spp
:il>ri-\
11.2
II
3 11.2
pnipiiriin
n
4 11.2
4
spp IISI-(I 11.2
spp
;il>ri-\
4
11.2
IT
I
4 11.2
pnipiiriin
n
1
11.3
spp
r 11 .3
spp IISI-(I
r Ti.3 r
spp 11.3
;il>l>ri-\ IT
I
1 I 1.3
pnipiiriin
n
I'. I'li-ii i\
i- I I I "
T:iri-
I llsli
III
iri'i'iinipiisi
ll-
1.00
0.25
1.55
1.56
2.43
1.41
0.10
Fish
Median all
fish eating
median all
invertebrat
es
f+a
2.28
0.54
2.00
1.60
3.20
1.46
0.15
Fish
Median all
fish eating
median all
invertebrat
f+a
2.28
0.15
2.06
1.29
2.66
1.48 1.01	1.49
Fish Median all f+a 2.28
fish eating
median all
invertebrat
es
1.74
0.15 TL1
TL1
1.00
0.15
0.8	2.25 0.86	1.93
1.91 0.71	1.36
B-169

-------
Ki-l'i-ri-n Sili-	Sili-
ii-	di-siripi II)
inn
T;iri-l
II sli
spi-iii-s
I'limmii
1111:11111-
l-'ish
III
I'isli |)n\ :is di-si rihi-d in \:iluri-Si-r\i-
i"ish pn-\ spp 1 11.2111 i ti.2 r rn.2 2 11.2
imiimi-nl	spi-iii-s usi-d spp	11.2 prupurliii spp usi-d
;il>l>ri-\ II n
">o
11.2
spp
;il>ri-\
I	1.2
II
I
2 I 1.2
pnipiiriin
n
3 11.2
spp iim-(I
I 1.2
spp
:il>ri-\
11.2
II
I
3 11.2
pnipiiriin
n
4 11.2
spp IISI-(I
4
11.2
spp
;il>ri-\
4
n.2
ii
i
4 11.2
pnipiiriin
n
1
11.3
spp
r 11 .3
spp IISI-(I
rn.3 r
spp 11.3
;il>l>ri-\ IT
I
1  TI.3	IH'li-iiiN T:iri- Tl Tiiimposi
pnipiiriin	i- III I llsli k-
n	III
Default
Default
Default
Default
Default
stonecat 0.77 Eats mainly bottom invertebrates (insects,
crayfish); sometimes also plant material
and fishes (Becker 1983, Scott and
Crossman 1973).
sucker 1.01 White: Larvae feed near surface on
protozoans, diatoms, small crustaceans,
and bloodworms. Adults feed
opportunistically on bottom organisms,
both plant and animal (e.g., chironomid
larvae, zooplankton, small crayfishes)
(Becker 1983, Sublette et al. 1990).
Bluehead: A bottom feeder. Scrapes
algae and other organisms from rocks
with chisel-like ridges inside each lip;
ingests fine organism-laden sediments.
May feed in stream riffles, or deeper
rocky pools; in lakes it may feed over
rocks near shore. May eat aquatic insect
larvae. Flannelmouth: Bottom feeder.
Reported to feed on diatoms, algae,
fragments of higher plants, seeds, and
benthic invertebrates (Sigler and Miller
1963; Lee et al. 1980). See Tyus and
Minckley 1988 for possible importance
of Mormon cricket as food source,
tilapia 1.41 aureus: Eats mainly phytoplankton.
mossambicus: Nonselective omnivore;
eats planktonic algae, aquatic plants,
invertebrates, and small fishes (Moyle
1976). zilli: Feeds on algae and higher
plants, invertebrates, and occasionally
eats dead or dying fish,
tui chub 1.20 Adults opportunistic. They feed on plant
material, plankton, insect larvae,
crustaceans, fish fry and fish eggs, etc.
Young feed on zooplankton. Coarse-
rakered form eats more plant material,
fme-rakered form more zooplankton.
Utah	1.01 Bottom feeder. Varied diet; feeds freely
sucker	on both animal and plant organisms, at all
depths throughout the year. Grazes on
filamentous algae.
Median
insects and
benthic
crustaceans
Median of all
invertebrates
except
bivalves
in,be-
all
1.74
1.48
0.70 TL1
0.50 TL1
TL1
TL1
1.00
1.00
0.20
0.50
l'ish Median all
fish eating
median all
invertebrat
f+a
2.28
0.1
1.65 0.77	1.26
1.24 1.01	1.25
Median of all
invertebrates
except
bivalves
TL1
all
1.48
0.50 TL1
TL1
1.00
0.50
1.24 1.41	1.74
TL1
1.00
0.40 Median of pc
planktonic
crustacean
1.41
Median of all
invertebrates
except
bivalves
all
1.48
0.50 TL1
TL1
1.00
0.28 Median of in
all insects
0.50
2.14
0.28 crayfish cr	1.46
0.04
1.45 1.20	1.75
1.24 1.01	1.25
B-170

-------
Ki-l'i-ri-n Sili-	Sili-
ii-	di-siripi II)
inn
T;iri-l I'isli
iisii	in
spi-iii-s
imiiinii
11 11:11111-
I'isli |)n\ :is di-si rihi-d in \:iluri-Si-r\i-
i"ish pn-\ spp 1 11.2111 i ti.2 r rn.2 2 11.2
imiimi-nl	spi-iii-s usi-d spp	11.2 prupurliii spp usi-d
;il>l>ri-\ II n
">o
11.2
spp
;il>ri-\
I	1.2
II
I
2 I 1.2
pnipiiriin
n
3 11.2
spp iim-(I
I 1.2
spp
:il>ri-\
11.2
II
I
3 11.2
pnipiiriin
n
4 11.2
spp IISI-(I
4
11.2
spp
;il>ri-\
4
n.2
ii
i
4 11.2
pnipiiriin
n
1
11.3
spp
r 11 .3
spp IISI-(I
rn.3 r
spp 11.3
;il>l>ri-\ IT
I
1  TI.3	IH'li-iiiN T:iri- Tl Tiiimposi
pnipiiriin	i- III I llsli k-
n	III
Default
walleye 1.60 Adults feed opportunistically on various
fishes and larger invertebrates.
Median
insects and
benthic
crustaceans
in,be
1.74
0.50
fish
Median all
fish eating
median all
invertebrat
f+a
2.28
0.5
2.01 1.60	3.21
Default
Default
Default
Default
western 1.21 Opportunistic omnivore; eats mainly
mosquit	small invertebrates, often taken near
ofish	water surface. Also eats small fishes and,
in the absence of abundant animal food,
algae and diatoms (Moyle 1976).
Mosquito fish are principally carnivorous,
and have strong, conical teeth and short
guts (Meffe et al. 1983, Turner and
Snelson 1984). They are reported to feed
on rotifers, snails, spiders, insect larvae,
crustaceans, algae, and fish fry, including
their own progeny (Barnickol 1941,
Minckley 1973, Meffe and Crump 1987).
Cannibalism has been documented by
several authors (Seale 1917, Krumholz
1948, Walters and Legner 1980,
Harrington and Harrington 1982). Plant
material is taken occasionally (Barnickol
1941) and may make up a significant
portion of the diet during periods of
scarcity of animal prey (Harrington and
Harrington 1982). Grubb (1972) showed
that anuran eggs from temporary ponds
were preferentially selected over those
breeding in permanent systems,
white	1.48 Eats fishes, zooplankton, aquatic insects,
bass	oligochaetes, and crayfish; fishes often
dominate diet of adults; diet may vary
from place to place (Moyle 1976,
Sublette et al. 1990).
white	2.67 Eats fishes, planktonic crustaceans, and
crappie	aquatic insects; small individuals eat
mostly zooplankton, fish tend to
predominate in the diet of larger
individuals, though zooplankton also
consumed (Moyle 1976)
white	1.21 A bottom feeder. Young feed mostly on
sturgeo	the larvae of aquatic insects, crustaceans,
n	and molluscs. A significant portion of the
Median of all in
insects
2.14
0.75
Median of
all
crustacean
s
1.41
0.25
1.96 1.21	2.37
Median of all in	2.14	0.30 Median of pc 1.41	0.05 crayfish cr	1.46	0.05
insects	planktonic
crustacean
s
Median of all in	2.14	0.50 Median of pc 1.41	0.10
insects	planktonic
crustacean
s
Median	in,be 1.74
insects and
benthic
0.31 Median of bvs 4.29	0.09
all bivalves
Fish Median all f+a 2.28
fish eating
median all
invertebrat
es
Fish Median all f+a 2.28
fish eating
median all
invertebrat
Fish Median all f+a 2.28
fish eating
median all
0.6
0.4
2.15 1.48	3.19
2.12 2.67	5.66
0.6
2.29 1.21	2.77
B-171

-------
Ki-li-mi Sill- Sill- T;iri-I	I isli I isli pn-> :is ili-si-rihi-il in Viliiii-Si-ni- I isli pn\ spp 1 11.2 I I I" 1 11.2	1	1 11.2	2 11.2 2	2	2 11.2	3 11.2 3	3	3 11.2	4 11.2 4	4	4 11.2	1	1 11.3 1 11.3	1	l TI.3	i:ili-ui\ T:iri- l l l'iiuiiposi
ii- di-siripi II) IIsli	III iiiiiiiiii-nl spi-iii-s usi-il spp	11.2	pnipiiriin	spp usi-d 11.2	TI.2	pniporliii	spp usi-il 11.2	TI.2	pn>pi>rlii>	spp usi-il 11.2	TI.2	pn>pi>rlii> TI.3	spp usi-il spp	T1.3	pniporiio	111"	i llsli k-
iiili spi-iii-s ;il>l>ri-\	II	11 spp	II	n spp	IT	n spp	IT	n	spp ;il>l>ri-\	II	II	III
ionium	I-"	;ihiv\ l"	:il>ri-\	I-"	;il>ri-\	l"	I-"
n n.imi-
dicl oi" laraer slur" con consists oi" fish.	crustaceans
in vertebral
es
Default
Default
white
sucker
wiper
1.11
1.48
Adults feed opportunistically on bottom
organisms, both plant and animal (e.g.,
chironomid larvae, zooplankton, small
crayfishes) (Becker 1983, Sublette et al.
1990).
Adults are predatory on fishes and larger
crustaceans (Hassler 1988).
expected
common spp in
benthos
TL1
crayfish
TL1
1.00
1.46
0.50 Median of
all insects
0.20
2.14
0.30
Median of
planktonic
crustacean
s
pc
1.41
0.10 crayfish cr
1.46
0.10
Fish Median all f+a
fish eating
median all
invertebrat
2.28
1.43 1.11
1.58
0.8	2.11 1.48	3.13
Default
yellow 1.42 Larvae and young primarily zooplankton
perch	feeders; older young eat mostly
invertebrates associated with bottom and
with aquatic plants; adults feed among
plants and along bottom on larger
invertebrates and small fishes (Moyle
1976).
Median
insects and
benthic
crustaceans
in,be 1.74
0.64 Median of pc
planktonic
crustacean
s
1.41
0.13 TL1
TL1 1.00
0.07
Fish Median all f+a
fish eating
median all
invertebrat
es
2.28
0.16	1.73 1.42	2.47
Saiki et
al. 1993
Saiki et
al. 1993
Saiki et
al. 1993
bluegill 1.03
largemo
uth bass
western
mosquit
ofish
site-specific: 0.23 chironomid; 0.3
microcrustacea; 0.47 amphipod
1.39 site-specific: 0.73 fish; 0.27 crayfish
1.21 site-specific: 0.89 molluscs, and insects;
0.065 chironomid; 0.045 microcrustacea
stomach
analysis
stomach
analysis
stomach
contents show
a large
terrestrial
component
amphipods am	1.22
crayfish
Median
insects and
benthic
crustaceans
1.46
in,be 1.74
0.47 Median of
planktonic
crustacean
s
0.27 Saiki
bluegill
TTFcomp.
0.89 midges
pc
1.41
BG 1.47
mg
1.90
0.30 midges mg 1.90
0.73
0.07 Median of pc 1.41
planktonic
crustacean
s
0.23
0.05
1.43 1.03	1.47
1.47 1.39	2.04
1.74 1.21	2.10
Formatio
Crow
CC-
brown
n 2012
Creek -
150
trout

CC150


Formatio
Crow
cc-
sculpin
n 2012
Creek -
150


CC150


1.38 Proportions as described in table B-10
1.29 Proportions as described in table B-10
Median of all in	2.14 0.57	midges mg 1.90 0.27	mayflies mf 2.38 0.16
insects
Median of all in	2.14 0.57	midges mg 1.90 0.27	mayflies mf 2.38 0.16
insects
2.12 1.38
2.12 1.29
2.91
2.74
B-172

-------
Ki-li-n-n
Sili-
Sik-
T;iri-I
ii-
di-siripi
ll)
11 sli

inn

spi-iii-s



iniiinni



n n.imi-
Formatio
Crow
i_i_-
bro\ui
n 2012
Creek -
1A
trout

1A


Formatio
Crow
CC-
sculpin
n 2012
Creek -
1A


1A


Formatio
Crow
CC-
brown
n 2012
Creek -
350
trout

CC350


Formatio
Crow
cc-
sculpin
n 2012
Creek -
350


CC350


Formatio
Crow
cc-
brown
n 2012
Creek -
3A
trout

3A


Formatio
Crow
cc-
sculpin
n 2012
Creek -
3A


3A


Formatio
Crow
cc-
brown
n 2012
Creek -
75
trout

CC75


Formatio
Crow
cc-
sculpin
n 2012
Creek -
75


CC75


Formatio
Deer
DC-
brown
n 2012
Creek
600
trout
Formatio
Deer
DC-
sculpin
n 2012
Creek
600

Formatio
Hoopes
HS
brown
n 2012
Spring -

trout

HS


Formatio
Hoopes
HS
sculpin
n 2012
Spring -



HS


Formatio
Hoopes
HS-3
brown
n 2012
Spring -

trout

HS3


Formatio
Hoopes
HS-3
sculpin
n 2012
Spring -



HS3


Ill
I'isli piv\ :is (k-siril)i'(l in VilinvSini-
i"ish |)iv\ s|>|> i 11.2111 i n.2 r	rn.2 2 11.2
imiimi-nl	spicks usi-il spp	11.2	pnipiiriiii spp usi-d
ill>l>IY\ II	11
I
2 2 2 11.2 3 11.2 3 3 3 11.2 4 11.2 4 4 4 11.2 1 1 11.3
TI.2 I I.2 pnipiiriiii spp usi-il I I.2 TI.2 pnipiiriiii spp usi-il I I.2 TI.2 pnipiiriiii TI.3 spp usi-il
n	spp TT n	spp
spp II
;il)iv\ I-"
ii
spp II
:il>ri-\ I-"
spp II
;il>ri-\ l"
rn.3 r	in.3
spp 11.3	priipnrlin
;il>l>ri-\ IT	n
I
KITiiliv T:iri- lll'iiimposi
III 1 fish ii-
III
1.38 Proportions as described in table Li-10
1.29 Proportions as described in table B-10
1.38 Proportions as described in table B-10
1.29 Proportions as described in table B-10
1.38 Proportions as described in table B-10
1.29 Proportions as described in table B-10
1.38
1.29
1.38
1.29
1.38
1.29
1.38
1.29
Proportions as described in table B-10
Proportions as described in table B-10
Proportions as described in table B-10
Proportions as described in table B-10
Proportions as described in table B-10
Proportions as described in table B-10
Proportions as described in table B-10
Proportions as described in table B-10
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all bvs
bivalves
Median of all bvs
bivalves
Median of all crs
crustaceans
Median of all crs
crustaceans
2.14 0.79	midges mg 1.90 0.09	mayflies mf 2.38 0.12
2.14 0.79	midges mg 1.90 0.09	mayflies mf 2.38 0.12
2.14 o.:
2.14 o.:
1.41 0.40
1.41 0.40
midges mg 1.90 0.07	mayflies mf 2.38 0.13
midges mg 1.90 0.07	mayflies mf 2.38 0.13
2.14 0.85	midges mg 1.90 0.05	mayflies mf 2.38 0.10
2.14 0.85	midges mg 1.90 0.05	mayflies mf 2.38 0.10
2.14	0.49	midges	mg	1.90	0.38
2.14	0.49	midges	mg	1.90	0.38
2.14	0.44	midges	mg	1.90	0.21
2.14	0.44	midges	mg	1.90	0.21
4.29	0.44	midges	mg	1.90	0.32
4.29	0.44	midges	mg	1.90	0.32
mayflies mf 2.38 0.13
mayflies mf 2.38 0.13
mayflies
mayflies
mf
mf
blackwor bw
ms
blackwor bw
ms
Median of in
all insects
Median of in
all insects
2.14 0.33
2.14 0.33
2.38	0.35
2.38	0.35
1.29	0.24
1.29	0.24
mayflies mf 2.38 0.27
mayflies mf 2.38 0.27
2.15 1.38
2.15 1.29
2.16 1.38
2.16 1.29
2.15 1.38
2.15 1.29
2.08 1.38
2.08 1.29
2.18
2.18
2.81
1.38
1.29
1.38
2.81 1.29
1.91 1.38
1.91 1.29
2.96
2.78
2.97
2.79
2.97
2.78
2.87
2.69
3.00
2.81
3.86
3.63
2.63
2.47
B-173

-------
Ki-li-mi Sili-
ii-	di-siripl
inn
Sili- T:iri-I I'isli
II)	fish'	III
spi-iii-s
imiiinii
11 11:11111-
I'isli |)n\ :is di-si rihi-d in \:iliiri-Si-r\i-
i-'ish pn-\ spp 1 11.2111 i ti.2 r	rn.2 2 11.2
imiimi-nl	spi-iii-s usi-d spp	11.2	pnipiiriio spp usi-d
;il>l>ri-\ II	11
I
2 2 2 11.2 3 11.2 3 3 3 11.2 4 11.2 4 4 4 11.2 1 1 11.3
TI.2 I I.2 pnipuriiii spp usi-d I I.2 TI.2 pnipiiriin spp usi-d I I.2 TI.2 pnipiiriin TI.3 spp usi-d
n	spp TT n	spp
spp II
;il)iv\ I-"
ii
spp II
:il>ri-\ I-'
spp II
;il>ri-\ I-'
rn.3 r	in.3
spp TI.3	pniporiid
;il)l)n-\ 'IT	n
I
IH'li-iiiN 'l':ni- I ri'i'iiiiiposi
i- I TI-" 1 llsli k-
III
r'onnalio
n 2012
Fonnatio
n 2012
Fonnatio
n 2012
Fonnatio
n 2012
Fonnatio
n 2012
Fonnatio
n 2012
Sage
Creek -
LSV2C
Sage
Creek -
LSV2C
Sage
Creek -
LSV4
Sage
Creek -
LSV4
South
Fork
Tincup
Cr.
South
Fork
Tincup
Cr.
LSV- brown
2C trout
LSV- sculpin
2C
LSV-
4
brown
trout
LSV-	sculpin
4
SFTC	brown
-1	trout
SFTC sculpin
-1
1.38 Proportions as described in table Li-10
1.29 Proportions as described in table B-10
1.38 Proportions as described in table B-10
1.29 Proportions as described in table B-10
1.38 Proportions as described in table B-10
1.29 Proportions as described in table B-10
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
Median of all in
insects
2.14 0.57	midges mg 1.90 0.12	mayilies ml" 2.38 0.31
2.14 0.57	midges mg 1.90 0.12	mayflies mf 2.38 0.31
2.14 0.53
2.14 0.53
2.14 0.93
2.14 0.93
midges mg 1.90 0.34
midges mg 1.90 0.34
mayflies mf 2.38 0.13
mayflies mf 2.38 0.13
Median of bvs 4.29 0.03	mayflies mf 2.38 0.04
all bivalves
Median of bvs
all bivalves
4.29 0.03	mayflies mf 2.38 0.04
2.19 1.38
2.19 1.29
2.09 1.38
2.09 1.29
2.22 1.38
2.22 1.29
3.01
2.83
2.70
3.05
B-174

-------
Table B-9. Calculation of site-specific invertebrate proportions using invertebrate counts in Formation 2012
Ol'tlrl
< Jrllll-
Ihiliilnl I'liiHlinii liilri'iiii Sliviiin SI- I iiHiip ( i vrk
ill
l!cki\i I ii-diim
( i'oh ( ivrk


nr




(Mil
Ephemeropt
Atenella
CN
CG
era
margarita


Ephemeropt
Baetis spp.
sw
CG
era



Ephemeropt
Centroptilum
sw
CG
era
conturbatum


Ephemeropt
Cinygmula
CN
sc
era
spp.


Ephemeropt
Diphetor
SW
CG
era
hageni


Ephemeropt
Drunella
CN
P
era
coloradensis


Ephemeropt
Drunella
CN
P
era
grandis


Ephemeropt
Epeorus
CN
SC
era
longimanus


Ephemeropt
Ephemerella
CN
CG
era
dorothea



infrequens


Ephemeropt
Ephemerella
CN
CG
era
aurivillii


Ephemeropt
Paraleptophle
SW
CG
era
bia spp.


Ephemeropt
Tricorythodes
CN
CG
era
minutus


Plecoptera
Hesperoperla
CN
P

pacifica


Plecoptera
Isoperla sp.
CN
P
Plecoptera
Malenka sp.
CN/SP
SH
Plecoptera
Pteronarcys
CN/SP
SH

sp.


Plecoptera
Skwala sp.
CN
P
Plecoptera
Sweltsa sp. p
CN
P
Trichoptera
Agapetus sp.
CN
SC
Trichoptera
Arctopsyche
CN
P

sp.


Trichoptera
Brachycentrus
CN
F

sp.


Trichoptera
Cheumatopsy
CN
F

che sp.


Trichoptera
Dicosmoecus
BU
SH

sp.


Trichoptera
Dolophilodes
CN
F
si- I ( I
( ( "5
< ( 1511
< < .550
I i x iil iii
ii
Dull' K 2') 2IMI" '>'>200 ') 2 2 .i 200S '>1200 S 24 200 > .i 200 ') I 200 S 2.5 200 ') 4 200
X	(i	~	(i	~	S	(i	~	s
21
10
3
7
12
12
33
21
56
14
5
9
11
25
14
4
4
85
7
16
18
29
89
11
11
27
90
14
35
38
2
61
7	7
3	4	7
2
13
l)i'i'i*( reck
llllll|ll"> *S|l|-ill>i
S;mr ( ivrk
('rim ( reck
IX (.1111
IIS
)" 200 S 2" 200 ')X 200X ')X200 X 24 200"
(i	~	(i
253
2
1
1
76
14
67
76
11
I is.?
I S\ 2(
I s\ 4
( ( I \
< ( ,5\
'> 4 200 ) (, 2000
X
56
X 2X 20 '> 5 200X ') X 200 X 2X 20 '> 5 200 ') 5 200 ) I 200 X 25 20 ') (. 200 ) 4 200 X 2(. 20
0"	(	0"	X	(i	(i	0"	X	(i	0"
249
316
27
53
46
57
32
1
3
62
56
) ~ 200
X
61
I nl:i
I
5
2041
1
16
30
29
38
12
25
11
21
14
13
62
30
35
14
13
7
20
14
2
23
13
17
33
65
34
153
23
29
1
29
20
73
11
27
13
61
3
15
18
25
12
75
25
217
46
212
24
37
77
2
191
635
21
7
28
B-175

-------
Orilrr
< Jrllll-
ILiliilnl
l!ih;i\ i
Ml
liiiHliiin
:il
1 llllTllll
IT
SIIVlllll
1 oi'iiliu
SI- 1 iiu ii|) ( irrk
SI 1(1

< < "5

( row ( ivrk
< '< 1511


< < .550




(ir<>ii|iN

II
Ifcilr
X 2') 2IHI" ')') 2IHI
S
') 2 21 III
(
S 2.? 21II1
').? 2IHIS
) 1 200 S 24 200
(.
').? 200
s
) 1 200
(
S 2.? 200
') 4 200
S

sp.













Trichoptera
Glossoma sp.
CN
SC
0










Trichoptera
Helicopsyche
sp.
Hesperophyla
x sp.
Hydropsyche
sp.
Hydroptila sp.
CN
sc
3


3


5 4

5
93

Trichoptera
Trichoptera
CN
CN
SH
F
5
4

5
47
50
23
29 17
11
74
97
41
Trichoptera
CN
SC
6


8
9
1
1




Trichoptera
Trichoptera
Lepidostoma
spp.
Micrasema sp.
SP/CB
CN
SH
SH
1
1

1 3
8
7
9
6 8
65 1
18
14
67
6
Trichoptera
Neothremma
sp.
Oecetis
disjuncta
Onocosmoecu
s sp.
Oligophlebod
es sp.
Parapsyche
sp.
Psychoglypha
sp.
Rhyacophila
spp.
Wormaldia
spp.
Ametor sp.
CN
SC
0










Trichoptera
Trichoptera
Trichoptera
Trichoptera
CN
CB
CN
CN
P
SH
SC
P
8
1
1
1

2
1
11
16


5

2
3

Trichoptera
SP/CB
CG
1



3






Trichoptera
Trichoptera
Coleoptera
CN
CN
sw
P
F
P
0
3
5

7 3
3
4
5
5
9 17
1 8
16
15
3
2
5
5
9
6
Coleoptera
Brychius sp.
CB
SC
7

2

1
3





Coleoptera
Cleptelmis sp.
CN
CG/SC
4

3 26







4
Coleoptera
Dubiraphia sp.
CN
CG/SC
4

3








Coleoptera
Coleoptera
Heterlimnius
corpulentus
Optioservus
quadrimaculat
us
Oreodytes sp.
CN/BU
CN
CG/SC
CG/SC
4
4

97 267
43
109
68
40 205
153
78
162
167
Coleoptera
SW/DV
P
5

6
1







Coleoptera
Paracymus sp.
CN
P/OM
5





1




Coleoptera
Megaloptera
Zaitzevia
paravula
Sialis sp.
CN/BU
BU/CB
CG/SC
P
4
4

170 57
1
5
4
1
3
1 1
1
7
23
5
18
I)rrl"< Irrk
ll(iii|H"> S|iiiim
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1	14
16
16
7
23
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83
32
5
28
12
11
3
21
48
13
9
33
16
10
18
2
14
9
132
4
11
151
53
76
63
1
27 153
19	81
91	29 105	79
2
13	4	2
3	36
74 246
16	11
69
83
151
11
4
214
70
1012
39
141
273
4
17
1
13
32
3
236
77
1
16
6 46
3
67
129
13
2556
7
1
366
B-176

-------
I
I
9
5
1
37
119
20
13
8
3
4
74
29
4
3
11
1
760
27
2168
1
2
405
69
174
10
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188
15
2
5
7
195
30
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173
26
143
2
6
2
49
17
99 143
1
17
1
102
10	30
33
15
3
151
19
92
72
2
52
124
25
101	5
5	2
57	27
13	21
3
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38
24
20
34
12
25
2
43
24
91
19
6
114
55
14
32
B-177

-------
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-------
Table B-10. Summary of Formation 2012 invertebrate data.
l'livlum
Subphylum
Class
Subclass
Inlraclass
Superorder
()rdcr
Lookup II)
Common name
SI
1 CI
( ( "
(









Counl
l'roporlion
Counl
l'roporlion
Connl
Arthropoda

Insecta
Pterygota

Ephemeropteroidea
Ephemeroptera
Ephemeroptera
Mayflies
36
0.04
185
0.12
231
Arthropoda

Insecta
Pterygota

Exopterygota
Plecoptera
Plecoptera
Stoneflies
107
0.12
85
0.06
40
Arthropoda

Insecta


Amphiesmenoptera
Trichoptera
Trichoptera
Caddisflies
30
0.03
268
0.18
283
Arthropoda

Insecta
Pterygota
Neoptera
Endopterygota
Coleoptera
Coleoptera
Beetles
631
0.72
234
0.15
408








Alderflies,













dobsonflies and





Arthropoda

Insecta

Neoptera

Megaloptera
Megaloptera
fishflies
1
0.00
4
0.00
1








Dragonflies and





Arthropoda

Insecta
Pterygota

Odonatoptera
Odonata
Odonata
damselflies













True bugs (cicadas,













aphids,













planthoppers,













leafhoppers, shield





Arthropoda

Insecta

Neoptera
Paraneoptera
Hemiptera
Hemiptera
bugs)
5
0.01



Arthropoda

Insecta


Panorpida
Diptera
Diptera
True flies
42
0.05
143
0.09
103






Chironomidae
Chironomidae






Arthropoda

Insecta



(family)
(family)
Midges


556
0.37
385







Hirudinea






Annelida

Clitellata
Hirudinea



(class)
Leeches




1








Springtails (not





Arthropoda

Entognatha



Collembola
Collembola
insects!)




2







Oligochaeta






Annelida

Clitellata
Oligochaeta



(class)
Worms
5
0.01
24
0.02
17







Bivalvia






Mollusca

Bivalvia




(class)
Clams
2
0.00
6
0.00
2







Gastropoda






Mollusca

Gastropoda




(class)
Snails and slugs
21
0.02
7
0.00

Arthropoda
Crustacea
Malacostraca



Amphipoda
Amphipoda
Crustaceans





Arthropoda
Crustacea
Ostracoda




Ostracoda
Sea shrimp




1
Platyhelminthes
Turbellaria



Tricladida
Tricladida
Flatworms












Acari






Arthropoda
Chelicerata
Arachnida
Acari



(subclass)
Mites and ticks


4
0.00
7







Total

880

1516

1481








Midge

0.00

0.37









Mayfly

0.04

0.12









Other insects

0.93

0.48









Molluscs

0.03

0.01









Crustaceans

0.00

0.00









Annelids

0.01

0.02









Other

0.00

0.00

50	CC350	IX'600	IIS
roporlion Counl Proportion Counl Proportion Counl	l'roporlion Counl
192
18
539
457
0.13
0.01
0.36
0.30
444
195
229
76
0.34
0.15
0.18
0.06
115
59
34
18
IIS3	I.SV2C	I.SY4	CC1A	CC3A
l'roporlion Counl l'roporlion Connl l'roporlion Connl l'roporlion Connl Proportion
0.11
0.06
0.03
0.02
325
20
230
65
0.24
0.01
0.17
0.05
421
26
324
327
0.28
0.02
0.21
0.22
56
30
135
29
0.13
0.07
0.30
0.07
168
11
325
507
0.12
0.01
0.23
0.36
136
22
623
311
0.00
145
108
0.10
0.07
55
272
0.04
0.21
29
241
0.03
0.24
89
106
0.07
0.08
85
154
0.06
0.10
36
149
0.08
0.33
221
127
0.16
0.09
166
70
16
11
4
15
0.01
0.01
0.00
0.01
27
0.02
0.01
0.00
178
9
319
2
0.18
0.01
0.32
0.00
3
2
16
19
471
0.00
0.00
0.01
0.01
0.35
43
12
54
21
51
0.03
0.01
0.04
0.01
0.03
0.02
0.00
0.00
9
4
21
2
1
0.01
0.00
0.01
0.00
0.00
0.00
75
24
29
1505
1307
1004
1346
1518
445
1402
1468
0.26
0.16
0.56
0.00
0.00
0.01
0.01
0.07
0.13
0.77
0.01
0.00
0.01
0.01
0.21
0.34
0.42
0.01
0.00
0.02
0.00
0.24
0.11
0.14
0.33
0.00
0.18
0.00
0.08
0.24
0.30
0.01
0.36
0.00
0.00
0.10
0.28
0.50
0.04
0.05
0.03
0.00
0.33
0.13
0.52
0.00
0.00
0.02
0.00
0.09
0.12
0.76
0.02
0.00
0.01
0.00
0.05
0.09
0.77
0.04
0.00
0.05
0.00
B-179

-------
Total
1.00
1.00
Take the top 3
that are above
1%
Insects
Molluscs
Mayfly
0.93	Insects
0.03	Midge
0.04	Mayfly
1.00
0.50	Insects
0.38	Midge
0.13	Mayfly
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
1.00
0.58	Insects
0.27	Midge
0.16	Mayfly
1.00
0.79	Insects
0.07	Midge
0.13	Mayfly
1.00
0.44 Midge
0.21 Molluscs
Worms
and
0.35 leeches
1.00
0.32 Insects
0.44 Crustaceans
0.24 Mayfly
1.00
0.33	Insects
0.40	Midge
0.27	Mayfly
1.00
0.57	Insects
0.12	Midge
0.31	Mayfly
1.00
0.53	Insects
0.34	Midge
0.13	Mayfly
1.00
0.78	Insects
0.09	Midge
0.12	Mayfly
1.00
0.85
0.05
0.10
1.00
B-180

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APPENDIX C: Summaries of Chronic Studies
Considered for Criteria Derivation
White sturgeon C-2
Sacramento splittail C-12
Fathead minnow C-15
Flannelmouth & razorback suckers C-22
Northern pike C-24
Chinook salmon C-27
Rainbow trout & brook trout C-32
Cutthroat trout C-51
Dolly Varden C-65
Brown trout C-68
Desert pupfish C-86
Eastern and western mosquitofish C-103
Striped bass C-105
Bluegill sunfish C-106
Largemouth bass C-147
See Appendix E for descriptions of other, less conclusive studies with:
Rainbow trout
Fathead minnow
Sacramento splittail
White sucker
See Appendix E for descriptions of invertebrate studies.
C-l

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Tashjian, D.H., S.J. The, A. Sogomoyan and S.S.O. Hung. 2006. Bioaccumulation and chronic toxicity
of dietary L-selenomethionine in juvenile white sturgeon (Acipenser transmontanus). Aquatic
T oxicol .79:401-409.
Test Organism:
Exposure Route:
White sturgeon (Acipenser trans montanus)
Dietary only
Seleno-L-methionine was added to an artificial diet consisting of vitamin-free
casein, wheat gluten, egg albumin, dextrin, vitamin mix, BTM-mineral mix,
cellulose, corn oil, cod liver oil, choline chloride and santoquin; the measured
dietary concentrations were 0.4, 9.6, 20.5, 41.7, 89.8, 191.1 mg Se/kg dw.
Test Duration:
8 weeks
Study Design:	25 juvenile white sturgeon were placed in each of 24 90-L tanks. Treatments
were randomly assigned to the 24 tanks resulting in 4 replicates per dietary
treatment. Four fish from each tank were sampled after 0, 4 and 8 weeks for
weight, length, liver weight, condition factors, hepatosomatic indices, hemocrit,
histopathology, and selenium measurement in liver, kidney, muscle and gill
tissues. 8 fish after 0 and 8 weeks were sampled for whole body selenium
measurement.
Effects Data:	Sturgeon survival did not differ significantly among treatment groups after the 8-
week exposure with a mean survival rate of 99 across all groups. Fish fed 41.7 to
191.1 mg Se/kg dw exhibited significant declines in body weight (see table). All
other endpoints measured were as sensitive or less sensitive to selenium in the
diet as body weight.
Mean (SE) white sturgeon moisture, lipid and whole body Se after 8-week exposure
Treatment
group
Moisture, % ww
Lipid, % ww
muscle Se, mg/kg dw
whole body Se, mg/kg dw
0.4
76.8 (0.5) b
9.5 (4) abc
8.2 (0.6) e
5.2 (0.4) c
9.6
77.0 (0.7) b
9.5 (0.9) abc
17.2 (0.7) d
11.8 (0.9) b
20.5
76.8 (0.3) b
10.1 (0.4) ab
22.9 (1.5) c
14.7 (0.8) b
41.7
77.3 (0.5) b
9.6 (0.7) abc
36.8 (1.8) b
22.5 (1.4) a
89.8
78.5 (0.3) ab
7.6 (0.4) bed
52.9 (3.2) a
34.4 (2.3) a
191.1
80.0 (0.4) a
6.1 (0.4) cd
54.8 (2.8) a
27.5 (4.4) a
C-2

-------
Mean (SE) white sturgeon body weight increase after 8-week exposure
Treatment
group
Body weight
increase (%)
muscle Se, mg/kg dw
whole body Se, mg/kg dw
0.4
282.9 (4.6) a
8.2 (0.6) e
5.2 (0.4) c
9.6
285.5 (9.9) a
17.2 (0.7) d
11.8 (0.9) b
20.5
277.7 (6.1) a
22.9 (1.5) c
14.7 (0.8) b
41.7
191.0 (12.6) b
36.8 (1.8) b
22.5 (1.4) a
89.8
106.5 (5.8) c
52.9 (3.2) a
34.4 (2.3) a
191.1
28.6 (3.6) d
54.8 (2.8) a
27.5 (4.4) a
Letters denote statistical groupings among treatments within each exposure period (p<0.05).
Chronic Value:	Using the logistic equation with a log transformation of the exposure
concentrations (TRAP program), the ECi0 and EC2o values for reduction in body
weight are 15.08 and 17.82 mg Se/kg dw whole body and 27.76 and 32.53 mg
Se/kg dw muscle tissue.
C-3

-------
White sturgeon (Tashjian et al 2006)
350
300
250
200
150
100
50
0
,9
.8
Log([Se] muscle, mg/kg dw)
Parameter Summary (Logistic Equation Regression Analysis)
Parameter
Guess
FinalEst
Std Error
95%LCL
95%UCL
LogXSO
1.6006
1.6303
0.0314
1.5304
1.7301
S
1.6574
2.938
0.925
-0.005
5.882
Y0
284.2
286,3
18 9
226.1
346.5

Effect Concentration Summary

% Effect
Xp Est
95%LCL
95%UCL
50.0
42.69
33.92
53.72
20.0
32.53
21.17
49.99
10.0
27.76
15.63
49.30
5.0
23.98
11.75
48.93
06/19/2009 10 32	MED Toxic Response Analysis Model, Version 1 03
C-4

-------
White sturgeon (Tashjian et al 2006)
350
300
250
1
 200
* 150
E 100
1.5
1.6
1.2
1.3
1.4
.8
,7
.8
.9
1.0
1.1
Log([Se] whole body, mg/kg dw)
Parameter Summary (Logistic Equation Regression Analysis)
Parameter
Guess
FinalEst
Std Error
95%LCL
95%UCL
LogX50
1.3403
1.3750
0.0643
1.1702
1.5797
S
2.283
2.794
1.908
-3.277
8.865
Y0
284.2
294.2
45.0
151.0
437.3
Effect Concentration Summary
% Effect
Xp Est
95%LCL
95%UCL
50.0
23.71
14.80
37.99
20.0
17.820
6.890
46.090
10.0
15.078
4,160
54.655
5.0
12.926
2.587
64.584
06/19/2Q09 10:42	MED Toxic Response Analysis Modei, Version 1.03
C-5

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Linville, R.G. 2006. Effects of Excess Selenium on the Health and Reproduction of White Sturgeon
(Acipenser transmontanus): Implications for San Francisco Bay-Delta. Dissertation. University of
California at Davis.
Test Organism:	White Sturgeon (Acipenser trans montanus)
Exposure Route: Dietary only
Selenium was added to the treatment in the form of selenized yeast. Selenized
yeast (2.2%; Selenomax, Ambi Inc.) was added to a commercial salmonid diet
and pelleted with fish oil. For the control diet, the selenized yeast mixture
contained 1.3% selenized yeast and 98.7 tortula yeast. Only selenized yeast was
added to the treatment diet. After pelleting, the diet was allowed to air dry on
drying racks.
Test Duration:	Females were fed 0.3% body weight/day the experimental diet for 6 months.
Study Design:	16 adult female white sturgeon (approximately 5 years old, mean weight and fork
length: 22.71 kg and 134.59 cm) were exposed in a freshwater flow through
system to either the control diet (8 females in one tank fed 1.4 mg/kg Se) or
treatment (8 females in a separate tank fed 34 mg/kg Se, Se from selenized yeast)
for 6 months. After the 6 month dietary exposure, females were induced to spawn
and fertilized with non-exposed male milt. Eggs were hatched in jars keeping
eggs from each female separate. For each progeny cohort, 3000 larvae were
randomly distributed into 3 reps for stage 40 (intestinal portion is void of yolk
material, but stomach is not differentiated and is filled with yolk) sampling and 3
reps for stage 45 (yolk sac absorbed, start exogenous feeding) sampling. Se and
biological measurements were made in each replicate.
Effects Data:	No Se effects were observed for length or weight of larvae. Effects were
determined for edema (Table 1), skeletal deformities (Table 2) and larval survival
(Table 4). Because the mortalities for each cohort were recorded up to the time
the sample was collected for abnormalities, a combined effects variable can be
the total proportion of hatched larvae which were both alive and without any
abnormalities at stage 45 (Table 4). This was calculated as PS(l-PA), where PS
is the proportion survival in the test chambers prior to sampling and PA is the
proportion of the sample of surviving larvae with abnormalities. Binomial
confidence limits are included in Table 4 for percent survival and percent
abnormalities for each cohort to visualize significant differences among data
points and between data points and fitted curves. Such confidence limits cannot
be directly calculated for the combined effects variable, for which confidence
limits were estimated by combining the lower and upper confidence limits of the
individual effects variables using the same equation as above (this slightly
overestimates the confidence limit range).
In Table 4, only cohort T2 is significantly different from the controls, based both
on larval survival and abnormalities. That this selenium effect is also supported
by the microinjection studies of Linville, which showed large abnormality
frequencies for egg Se injected with >15 mg/kg, but little or no effect at lower
concentrations (this is only supporting information because direct injection of a
C-6

-------
specific form of Se is not a complete surrogate for setting effect concentrations
for maternally transferred Se). For cohort T3, the data for abnormalities indicate
some effects, but cannot be considered a definite effect concentration due to a
combination of considerations - overlapping confidence limits with controls, no
increase in mortality, limited information on within-cohort variability, and, based
on egg concentrations, no effects for cohort T1 at a higher concentration.
ECio Calculations: The combined effects variable is plotted versus Se concentration in the eggs in
Figure 1. With only one definite partial effect, TRAP cannot be used to estimate
a curve. Instead, the interpolation protocol is applied between the last two points
based on specifying the highest no-effect concentration (HNOEC), 11.0 mg/kg,
to be the EC0 in the interpolation equation and specifying the upper control
plateau (Y0 in TRAP) to be average survival of the lower four points. The
resultant TRAP slope is 3.0 and the interpolated ECi0 is 15.6 mg/kg.
The egg ECio of 15.6 mg/kg is slightly lower than the value of 16.3 mg/kg in the
previous draft (Figure 3). The lower value was due to the inclusion of larval
survival with abnormalities in the endpoint and using interpolation between the
last two points rather than a TRAP model of the dataset.
Linville (2006) similarly calculated a 10% effective dose (ED10) of the combined
skeletal and edema data of 15.3 mg Se egg/kg dw using a logit regression.
Linville (2006) also noted statistically significant differences using a Tukey
Honest Significant Difference (HSD) test between Se and control treatments with
respect to both the incidence of Stage 45 skeletal and total deformities,
respectively, for the maternal transfer study. These author-reported results
support the evidence of an effect of selenium in white sturgeon similar to the
ECio of 15.6 mg Se/kg egg dw interpolated by TRAP.
The combined effects variable is plotted versus Se concentration in muscle in
Figure 2. Unlike for the egg concentration, the muscle concentration for cohort
T3, with a small but not significant effect, is greater than that for cohort Tl, with
no effect, so that TRAP can be used to estimate a curve, although only barely so.
This analysis was by tolerance distribution analysis with the log-triangular
model. The resultant TRAP estimates are 100% for the control value and 8.8 for
the EC0 (about 11% below the Tl concentration); the standard deviation is 0.14
log units, equivalent to a slope of 3.7. The ECio estimate is 11.9 mg/kg.
Chronic Value:	The chronic value for combined deformities and larval survival using egg Se is
an ECio of 15.6 mg egg/kg dw. The chronic value for this same endpoint in
muscle tissue is an ECi0 of 11.9 mg muscle/kg dw.
C-7

-------
Table 1. Edema deformities.

Control


Treatment




Larval Se

Edema
Larval Se

Cohort
Edema (%)
(mg/kg dw)
Cohort
(%)
(mg/kg dw)

C3
0.00 (1)
2.43
T1
0.00(1)
11.6
Stage 36
C4
0.00 (1)
1.69
T2
0.00(1)
18.4

C5
0.00 (1)
2.67
T3
6.67(1)
7.75
Stage 40
C4
0.00 (3)
1.8
T1
0.00 (3)
11.6

C5
0.00 (3)
2.88
T2
4.44 2.22 (3)
20.4




T3
1.67  1.67 (2)
7.22
Stage 45
C4
0.00 (3)
1.96
T1
0.00 (3)
12

C5
0.00 (3)
2.59
T2
15.56  1.11 (3)
19.4




T3
0.00 (2)
7.61
Table 2. Skeletal deformities.

Control


Treatment




Larval Se


Larval Se

Cohort
Skeletal (%)
(mg/kg dw)
Cohort
Skeletal (%)
(mg/kg dw)

C3
0.00 (1)
2.43
T1
0.00(1)
11.6
Stage 36
C4
0.00 (1)
1.69
T2
0.00(1)
18.4

C5
0.00 (1)
2.67
T3
10.00(1)
7.75
Stage 40
C4
1.11  1.11 (3)
1.8
T1
0.00 (3)
11.6

C5
1.11  1.11 (3)
2.88
T2
14.44  1.11 (3)
20.4




T3
8.33  1.67 (2)
7.22
Stage 45
C4
0.00 (3)
1.96
T1
0.00 (3)
12

C5
0.00 (3)
2.59
T2
21.11  1.11 (3)
19.4




T3
13.33 3.33 (2)
7.61
C-8

-------
Table 3. Combined edema and skeletal deformities.
Control
Cohort
Affected (%)
Egg Se
(mg/kg)
Larval Se
(mg/kg)
Treatment
Abnormal
Cohort (%)
Egg Se
(mg/kg)
Larval Se
(mg/kg)
Stage 36
C3
0.00 (1)
2.46
2.43
T1
0.00 (1)
11
11.6

C4
0.00 (1)
1.61
1.69
T2
0.00 (1)
20.5
18.4

C5
0.00 (1)
2.68
2.67
T3
16.67 (1)
7.61
7.75


1.11 1.11






Stage 40
C4
(3)
1.61
1.8
T1
0.00 (3)
11
11.6


1.11 1.11



18.89  1.11



C5
(3)
2.68
2.88
T2
(3)
20.5
20.4





T3
10.00  0 (2)
7.61
7.22
Stage 45
C4
0.00 (3)
1.61
1.96
T1
0.00 (3)
11
12






27.78 2.94



C5
0.00 (3)
2.68
2.59
T2
(3)
20.5
19.4






13.33 3.33







T3
(2)
7.61
7.61
Table 4. Stage 45 data combined abnormalities and percent larval survival.
Cohort
Egg Se
(mg/kg)
Muscle Se
(mg/kg)
% Survival
(95% Binomial CL)
% Abnormal
(95% Binomial CL)
[# Abnormal]1
% Alive & w/o
Abnormalities
(95% Binomial CL)
C4
1.61
1.22
99.7
(98.9-99.9)
0.0
(0.0-4.2)
[0,0,0]
99.7
(95.7-99.9)
C5
2.68
1.48
99.7
(98.9-99.9)
0.0
(0.0-4.2)
[0,0,0]
99.7
(95.7-99.9)
T3
7.61
11.1
>99.6
(98.7-99.8)
13.3
(3.7-24.6)
[3,5]
86.4
(74.4-96.3)
T1
11
9.93
>99.6
(98.7-99.8)
0.0
(0.0-4.2)
[0,0,0]
99.7
(95.7-99.9)
T2
20.5
15.3
91.6
(90.1-92.8)
27.8
(18.8-38.3)
[7,8,10]
66.2
(55.6-75.4)
Bracketed numbers denote abnorma
larvae in each of the 2-3 replicates of n=30.
C-9

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(/)
 100 -
05
E
o
c
_Q
<
80 
60 
T*\
I
3
40 
20 
0
CL
2 3	5	10
mg Se/kg egg (dw)
20 30
Figure 1. White sturgeon percent alive and without abnormalities as a function of the logarithm of
selenium concentrations in eggs. TRAP is used to interpolate between the last two points; EC10 = 15.6
mg Se/kg egg dw.
(/)
g 100 +
"to
E
_Q
<
+->
D
O
o3


2 3	5	10
mg Se/kg muscle (dw)
20 30
Figure 2. White sturgeon percent alive and without abnormalities as a function of the logarithm of
selenium concentrations in female muscle. TRAP tolerance distribution analysis with the log-triangular
model; EC10 =11.9 mg Se/kg muscle dw.
C-10

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1.2
1.0
.8
.6
CL
.4
.2
0
0
Log(mg Se/kg egg dw)
Parameter	Initial Final
LogX50	1.416 1.4215
S	2 2.6306
Y0	0.96667 0.96667
Figure 3. TRAP analysis from previous draft. Initial estimate for slope set equal to or less than 2.645 (set
to 2 for this figure). ECi0 =16.3 mg/kg.
C-ll

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Teh, S.J., X. Deng, D-F Deng, F-C Teh, S.S.O. Hung, T.W. Fan, J. Liu, R.M. Higasi. 2004. Chronic
effects of dietary selenium on juvenile Sacramento splittail (Pogonichthys macrolepidotus). Environ. Sci.
Technol. 38: 6085-6593.
Test Organism:	Sacramento splittail (Pogonichthys macrolepidotus)', juveniles 7-mos.old
Exposure Route: Dietary only
Dietary Treatments: 8 graded levels of dietary Se; dietary levels obtained by combining selenized
yeast with Torula (non-active) yeast. Selenized yeast contained approximately
21% of Se as selenomethionine and proteinaceous Se forms. Diet was formulated
as pellets by mixing dry ingredients with water and oil, fan-dried, crumbled and
sieved. Analyzed levels: 0.4 (no selenized yeast), 0.7, 1.4, 2.7, 6.6, 12.6, and 57.6
mg/kg.
Fish were fed twice daily with a daily feeding rate of 3% BW in first 5 months
and then adjusted to 2% BW thereafter.
Test Duration:	9 months
Study Design:	A flow-through system with 40 fish/tank (24 total tanks) was used; each tank
held 90 L. Flow rate was 4 L/min. Water temperature was maintained at 23C for
6 months and then 18C for last 3 months due to failure of water heating system.
5 fish were sampled from each tank at 5 and 9 months and measured for gross
deformities, length, weight, Se in liver and muscle. Sections of the liver were
kept for histopathology. Condition factor (100 x BW/length), heptatosomatic
index (100 x liver weight/BW), BCF (total organ Se/dietary Se) were determined.
Effects Data:	Mortality was observed in the two highest dietary treatments: 10 and 34.3%,
respectively. No mortalities were observed in fish fed diets # 12.6 mg/kg. No
significant difference in growth of fish fed 12.6 mg/kg Se in diet, but there was in
the fish fed 26.6 mg/kg Se. See table below for levels of Se in fish at 9 months
and associated effects.
Authors determined prevalence of deformities was higher in fish fed 6.6 and 12.6
mg/kg Se in their diet, however a dose-response relationship did not occur (e.g.,
no deformities in high concentration). Gross pathology was a more sensitive
endpoint than growth.
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Summary of effects and assoc. dietary and tissue concentrations in Sacramento splittail
after 9 month exp.
Dietary conc'n mg/kg
0.4
0.7
1.4
2.7
6.6
12.6
26.0
57.6
Se in liver, mg/kg dw
20.1
18.6
20.0
23.0
26.8
31.3
40.4
73.7
Se in muscle, mg/kg dw
6.6
6.9
9.2
10.1
15.1
18.9
29.4
38.7
Liver histopathology (mean lesions scores, N=15)
Macrophage aggregate
0.13
0.07
0.2
0.27
0.40
0.20
0.20
0.85
Glycogen depletion
0
0
0.2
0
0.4
0.2
0
1.38
Single cell necrosis
0
0
0
0.07
0.13
0
0.07
0.46
Fatty vacuolar degeneration
0
0
0
0.2
0.53
0.07
0.2
0.08
Eosinophilic protein droplets
0
0
0
0
0
0
0.07
0.85
Sum of mean lesion scores
0.13
0.07
0.4
0.54
1.46
0.47
0.54
3.62
Gross Pathology (No. of deformities, N=15)
Facial deformities (eye, jaw, and mouth)
0
1
0
1
5
3
0
0
Body deformities (kyphosis, lordosis,
scoliosis)
0
0
4
2
3
1
1
0
Prevalence of deformity (%)
0
6.7
26.7
20
53.3
26.7
6.7
0
Chronic Value:	Using gross pathology as the endpoint (prevalence of deformities, %), the
NOAEC is 10.1 mg Se/kg dw and the LOAEC is 15.1 mg/kg Se dw in muscle
tissue; MATC or CV = 12.34 mg/kg Se in muscle dw.
The above concentrations in juvenile muscle tissue cannot be exactly translated
into an equivalent egg-ovary or whole-body concentration in adult splittail. But
using the median egg-ovary to muscle ratio of 1.59 for the family Cyprinidae, the
NOEC and MATC would represent 16.1 and 19.6 mg Se/kg egg-ovary. Using the
median muscle to whole-body ratio of 1.26 for the family Cyprinidae, the NOEC
and MATC would represent 8.04 and 9.83 mg Se/kg whole body. However,
appropriateness of these conversion estimates rests upon uncertain assumptions
that the muscle concentrations in juvenile splittails equal those of adult splittails
under the same exposure conditions, and that splittail tissue ratios are those
typical of the family Cyprinidae.
Comments:	The authors observed deformities including spinal deformities using fish that
were 7-months-old at test initiation. This is the only study in which deformities
were observed in fish that were not exposed maternally.
Deng et al. (2008) exposed Sacramento splittail juveniles (21-day post hatch) to
dietary selenium and dietary methylmercury in a two factorial design for four
weeks. No adverse effects (growth, condition factor, lethargy or abnormalities)
were observed in the selenium only exposures. The splittail accumulated
approximately 3.5 mg Se/kg ww muscle in the highest dietary exposure (35 mg
C-13

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Se/kg. Using the average percent moisture in fish muscle of 78.4% (May et al.
2000), the dw Se concentration is 16.2 mg Se/kg muscle indicating the
recommended CV does not over-estimate an effect concentration.
Rigby et al. (2010) re-analyzed the juvenile Sacramento splittail data generated in
the Teh et al. (2004) study. The authors used logistic regression to estimate EC
values for deformities on a culled data set which eliminated the three highest
dietary treatments due to their departure from a standard concentration-response
relationship. The ECi0 value for the culled data set was 7.9 mg Se/kg dw muscle
which is lower than the recommended CV of 12.3 mg Se/kg dw muscle. Due to
the lack of a concentration-response relationship across the entire dietary range
and the lack of effects in the Deng et al. (2008) study, an ECi0 of 7.9 mg Se/kg
dw muscle is too uncertain for a recommended CV. Although the recommended
CV of 12.3 mg Se/kg dw muscle is based on deformities (an uncertain response),
it is considered representative of an effect level for this species because of the
significant reductions in growth at the two highest test concentrations.
C-14

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Bennett, William N., Arthur S. Brooks, and Martin E. Boraas. 1986. Selenium uptake and transfer in
an aquatic food chain and its effects on fathead minnow larvae. Arch. Environ. Contam. Toxicol. 15:513-
517.
Fathead minnow (Pimephalespromelas; 2 to 8 day-old larvae).
Dietary only
Green alga, Chlorellapyrenoidosa were exposed to Se (H275Se04) in culture
water for 3 days. Rotifers, Brachionus calyciflorus, were cultured in chambers
with selenium containing green algae at the ratio of 25 jug algae/ml to 50 jug
rotifer/ml for 5 hr. The rotifers were filtered to separate them from the algae and
immediately heat-killed. The Se concentration in the rotifers was measured for
75Se activity.
9 to 30 days
Selenium uptake by larval fathead minnows was measured in three experiments.
Se-contaminated and control rotifers for feeding to larval fish were prepared in
advance using the low algae:rotifer ratio. Daily equal volumes of rotifers were
divided among five 800 mL polypropylene larval chambers. Three chambers
received Se-contaminated rotifers and two received control rotifers. The rotifers
were dead at the time of feeding (heat killed).
Larval fish were hatched from eggs spawned in the laboratory. After hatching,
active larvae were divided equally among the larval test chambers (daily renewal
exposures using dechlorinated Lake Michigan water). Larvae were initially fed
rotifers raised on control algae (no selenium). The age of the larvae when first
fed Se-contaminated rotifers was 4, 9, and 3 days post-hatch for experiments 1, 2,
and 3, respectively. Larval fish were fed Se-contaminated rotifers for 7, 9, and 7
days in the 3 experiments. A post-exposure observation period of 19 and 2 days
was used for experiments 1 and 2, respectively. During this time the larvae were
fed control rotifers. Daily, larvae from a replicate were removed from the test
chamber, washed, placed in a 20 ml vial, and counted for 75Se activity for 20 min.
All larvae were then placed in test chambers with fresh food rations. At the end
of the study all fish were individually dried and weighed.

Experiment 1
Experiment 2
Experiment 3
Initial feeding of control diet
(days)
3
8
2
Day Se diet first fed
4
9
3
Day Se diet last fed
11
17
9
Observation days on control diet
19
2
0
Age at study termination (days)
30
19
9
C-15
Test Organism:
Exposure Route:
Test Duration:
Study Design:

-------
Effects Data:

Experiment 1
Experiment 2
Experiment 3
Mean food Se concentration
(mg/kg)
>70
68
55
Food intake ((_ig rotifers/larva)
50
1330
1190
Initial larvae mean dry wt. at start
of Se-laden food (jug)
90
400
100
Final larvae mean dry wt. (jug) at
end of test
1470 (Control)
800 (Treatment)3
1888 (Control)
1354 (Treatment)3
475 (Control)
416 (Treatment)
Final mean larval Se content (jug
Se/larva)b
0.0062
0.0700
0.0248
Final mean larval Se
concentrations (mg Se/kg dw)
43.0
51.7
61.1
a Significantly different from the control.
b Values when Se-laden feeding was ended.
Selenium was measured in the test water during the feeding exposures, but the
concentrations were insignificant (0.84 |Jg/L). Survival was not affected by the
selenium exposures. Preliminary tests showed that fathead minnow larvae would
reach plateau concentrations of selenium within the 7- to 9-day exposure periods.
The food supply was sufficient to sustain growth of the larvae during the study,
according to the authors. The authors state that selenium uptake and higher
selenium content in experiment 2 larvae was due to their larger size and ability to
consume more rotifers/unit time. Se-exposed larvae were significantly smaller
(p<0.05) in mass than controls for experiments 1 and 2.
Chronic Value:	GM of mean larval Se concentrations measured in the three experiments, i.e.,
43.0, 51.7, and 61.1 mg/kg dw WB, respectively, is 51.40 mg Se/kg dw.
C-16

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Dobbs, M.G., D.S. Cherry, and J. Cairns, Jr. 1996. Toxicity and bioaccumulation of selenium to a
three-trophic level food chain. Environ. Toxicol. Chem. 15:340-347.
Rotifer (Brachionus calyciflorus), and fathead minnow (Pimephales promelas)
12 to 24 hr-old at start.
Dietary and waterborne
Water
Filtered and sterilized natural creek water supplemented with nutrients (Modified
Guillard's Woods Hole Marine Biological Laboratory algal culture medium) for
algal growth. Sodium selenate (Na2Se04) was added to test water to obtain
nominal concentrations of 100, 200, or 400 (.ig Se/L. Concentrations remained
stable and equal in each trophic level.
Control Diet
No selenium was added to the water medium for the alga; green alga was free of
selenium for the rotifer; and rotifers were free of selenium for the fathead
minnow.
Selenium Diet
Sodium selenate was added to the culture medium for the alga; green alga
thereby contained a body burden for the rotifer; and rotifers thereby contained a
body burden for the fathead minnow.
Dietary Treatments: Each trophic level had a different treatment. The green alga was exposed directly
from the water (1, 108.1, 204.9, 397.6 (.ig total Se/L); rotifers were exposed from
the water (1, 108.1, 204.9, 393.0 |_ig total Se/L) and the green alga as food (2.5,
33, 40, 50 mg Se/kg dry wt.); and the fathead minnow were exposed from water
(1, 108.1, 204.9, 393.0 (ig total Se/L) and the rotifer as food (2.5, 47, 53, 60 mg
Se/kg dry wt.).
Test Duration:	25 days
Study Design:	A flow-through system utilizing a stock solution of filtered and sterilized creek
water controlled at 2^C was used to expose three trophic levels of organisms.
Approximately one liter of media was pumped from the algal chamber into the
rotifer chamber each day. A cell density between 3 and 6 / 10' cells/ml was
delivered to the rotifer chambers. Rotifers were started at a density of 151.4  7.7
females/ml and one liter/day of rotifers containing culture water was
intermittently pumped into the minnow chamber. (B. calyciflorus has a life span
of about 7 days at 25C.) The pump was necessary to overcome the swimming
ability of rotifers to avoid an overflow tube. Larval fathead minnows
(35/chamber) were prevented from escaping by a screened overflow. Chambers
were cleaned daily and aeration was provided. All chambers were duplicated for
test replication and water was measured for selenium on days 0, 2, 6, 7, 11, 14,
17, 20, and 24. All algal and rotifer biomass and selenium samples were made on
these days. Fathead minnow chambers were measured for biomass, dissolved
selenium, and tissue selenium concentrations of days 0, 7, 11, 14, 20, and 24.
Test Organism:
Exposure Route:
C-17

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Additional measurements were made in the 200 (.ig Se/L test chambers on the
fathead minnow on day 16. Selenium concentrations were maintained near the
nominal concentrations and the standard deviation of mean concentrations was
less than 4 percent.
Effects Data:	Rotifers. Rotifers did not grow well and demonstrated reduced survival at all
selenium exposure concentrations during the 25 day test. By test day 7 only the
lowest test concentration (108.1 (ig/L) had surviving rotifers which showed a
decrease in selenium content from test days 18 through 25. A reduction in rotifer
biomass was discernable by test day 4 in the selenium treatments and since all
test concentrations had viable rotifer populations present, the effect level was
calculated using these data.
Effect of Dietary and Waterborne Selenium on Rotifers after 4 Days Exposure
Se in water, jig/L
Se in diet, mg/kg dw
Se in rotifer tissue,
mg/kg dw
rotifer biomass, mg/ml
dw
1
2.5
2.5
0.028
108.1
33
40
0.025
202.4
40
54
0.011
393
50
75
0.003
Fathead minnows. Due to the reduction of rotifer biomass in the higher test
concentrations, fish mortality and reduction in fish growth observed in the latter
days of the test was difficult to discern between effects from starvation and
selenium toxicity. The data from test day 8 was selected for determining the
effect of selenium on fathead minnows because starvation could be excluded as a
variable.
C-18

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Effect of Dietary and Waterborne Selenium on Larval Fathead Minnows after 8 Days Exposure
Se in water, jig/L
Se in diet, mg/kg dw
Se in fathead minnow
tissue, mg/kg dw
Average fish weight,
mg dw
1
2.5
2.5
0.8
108.1
47
45
0.7
202.4
53
75
0.4
393
60
73
0.2
Chronic Value:
Rotifers	42.36 mg Se/kg dw (EC2o)
Fish	< 73 mg Se/kg dw (LOAEC) - not amenable to statistical treatment; the LOAEC
was based on the observation that a >50 percent reduction in mean fish weight
occurred at this tissue concentration.
C-19

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Schultz, R. and R. Hermanutz. 1990. Transfer of toxic concentrations of selenium from parent to
progeny in the fathead minnow (Pimephalespromelas). Bull. Environ. Contam. Toxicol. 45:568-573.
Test Organism:	Fathead minnow (Pimephales promelas; Adults)
Exposure Route: Dietary and waterborne
Selenite was added to artificial streams which entered the food web; thus, fish
were also exposed to selenium in the diet.
Study Design:	Four Monticello artificial streams were used for the study which lasted from
September 1987 to September 1988. For each study, two streams (treated) were
dosed continuously to achieve 10 |a,g/L and two streams served as controls. Mean
selenium concentrations at the head of the treated streams were 9.8 1.2 and
10.3 1.7 |a,g/L, respectively. The concentrations of selenium measured in the
water from controls streams were all less than the detection limit, i.e., 2 (ig/L.
Spawning platforms were submerged into each stream. One subset of six embryo
samples (n = 2000 embryos per sample) were collected from the streams for
selenium analysis. Another subset of ten embryo samples were reared in
incubation cups receiving the same stream water dosed with sodium selenite via a
proportional diluter. The treated embryos in egg cups received an average 9.7 
2.6 |ag Se/L. Samples of hatched larvae were analyzed for selenium content while
others were inspected for occurrence of edema and lordosis. Prior to test
termination, female parents were seined. The mean selenium content in the
ovaries of seven to eight females from the treated and control streams was
reported.
Effects Data:	Edema and lordosis occurred in approximately 25 percent of the fish spawned
and reared in 10 (ig Se/L. Corresponding occurrence in control fish incubated in
the egg cups was only 1 and 6 percent, respectively. Table 1 provides the
abnormality observations and the selenium residues in the embryos and ovaries
from the control and treated streams. Although a case can be made that the Se
treatment had a higher rate of edema and lordosis, there are some problems that
add uncertainty to the estimation of an effect concentration (R. Erickson, pers.
comm.). Heavy mortality/loss of embryo/larvae during monitoring and the erratic
occurrence of the abnormalities (e.g., there is a significant incidence of edema in
only 3 of 10 replicates for the Se treatment) led to the conclusion that results
should not be used for criterion derivation. However, the data from this study
support the range of reproductive effect levels determined in other studies. The
Se concentration in embryos from the 10 (ig/L treatment stream of 3.91 mg/kg
ww converts to 25.6 mg/kg dw using 15.3% dw (N=3 range 14.7 - 15.6%) for
fathead minnow eggs (R. Erickson, pers. comm). The previous draft used the Se
concentrations in the ovaries collected at the end of the study for the effect
concentration estimate. However, it was determined that the embryos are a more
direct representation of Se exposure and toxicity to the larvae.
Chronic Value:	The LOEC for embryos is <25.6 mg Se/kg dw.
C-20

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Table 1. Percent Abnormalities in Fathead Minnow Larvae and the Associated Selenium Concentrations
in Embryos and Ovaries.	
Treatment
[Se] embryos,
mg/kg ww (SD)
[Se] ovaries,
mg/kg ww (SD)
Edema, % (SD)
Lordosis, % (SD)
Control
0.31 (0.01)
0.77 (0.14)
0.9 (2.2)
5.6(8.8)
10 jig/L
3.91 (1.87)
5.89 (2.21)
24.6(36.1)
23.4 (20.8)
SD = standard deviation
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Beyers, D.W. and Sodergren, C. 2001a. Evaluation of interspecific sensitivity to selenium exposure:
Larval razorback sucker versus flannelmouth sucker. Larval Fish Laboratory. Department of Fishery and
Wildlife Biology, Colorado State University, Fort Collins, Colorado.
Test Organism:
Exposure Route:
Study Design:
Effects Data
Chronic Value:
Larval flannelmouth sucker (Catostomus latipinnis) and larval razorback sucker
(Xyrauchen texanus)
Dietary and waterborne - laboratory exposure (28-d early life stage)
Continuous flow diluter supplied a range of aqueous test concentrations <1, 25.4,
50.6, 98.9, and 190.6 |_ig/L selenate. Well water was used as the dilution water.
Across the range of aqueous exposure concentrations, each test chamber was fed
the same daily ration of living rotifers containing selenium at <0.702, 1.35, 2.02,
4.63, and 8.24 mg/kg dw, respectively. Rotifers accumulated selenium from
algae (Chlorella vulgaris) exposed to 0, 25, 50, 100, and 200 :g/L selenate.
Replicated (n=4) exposure beakers using a randomized, balanced 5x2 factorial
design (1st factor - selenium; 2nd factor - species). Survival was monitored daily
and growth measured at the end of the 28-day exposure. Selenium was measured
in the larvae at the end of the 28-day exposure.
No survival effects were observed and there were no decreases in fish weight or
length. Fish mass was found to increase as a function of selenium concentration.
The chronic values for the flannelmouth sucker and razorback sucker were >10.2
and >12.9 mg Se/kg dw, respectively, based on the concentrations of selenium
measured in whole-body tissue of larval fish at the highest water and dietary
selenium concentrations.
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Beyers, D.W. and Sodergren, C. 2001b. Assessment of exposure of larval razorback sucker to selenium
in natural waters and evaluation of laboratory-based predictions. Larval Fish Laboratory. Department of
Fishery and Wildlife Biology, Colorado State University, Fort Collins, Colorado.
Test Organism:	Larval razorback sucker (Xyrauchen texanus)
Exposure Route: Dietary and waterborne - laboratory exposure (28-d early life stage)
Larvae were exposed in a daily static-renewal system to control water
(reconstituted very hard) and site waters: De Beque, Orchard Mesa, North Pond
diluted 50%, and North Pond. Each water type received either a control diet
(rotifers) or a diet previously exposed to the site water (site food: rotifers fed
algae exposed to respective site water).
Study Design:	Replicated (n=4) exposure beakers using a randomized, balanced 5x2 factorial
design (1st factor - test water type; 2nd factor - rotifers cultured in control water or
in site water). Survival was monitored daily and growth measured at the end of
the 28-day exposure. Selenium was measured in the larvae at the end of the 28-
day exposure.
Effects Data:	No survival effects were observed. There were no significant decreases in growth
of fish exposed to both site water and site food compared to fish exposed to
control water and control food. There was a significant increase in growth of fish
exposed to site water and control food relative to fish exposed to control water
and control food (p<0.0001). There were reductions in the growth of fish (14%)
exposed to site water and site food compared to site water and control food
(p<0.0001). Due to the lack of a dose-response relationship in both the
concentration of selenium in the food (rotifers) and growth, and the concentration
of selenium in the fish larvae and growth, the authors did not attribute the effect
of site food on the growth of fish to selenium.
Chronic Value:	The NOAEC for the razorback sucker larvae in the four site water types based on
selenium in whole-body tissue were: De Beque >5.45 mg Se/kg dw; Orchard
Mesa >11 mg Se/kg dw; North Pond 50% dilution >41.1 mg Se/kg dw; North
Pond >42 mg Se/kg dw. Because no significant effects were observed in larvae
exposed to North Pond water at >42 mg Se/kg dw whole-body tissue, this value
was selected as the chronic value for the study.
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Muscatello, J.R., P.M. Bennett, K.T. Himbeault, A.M. Belknap and D.M. Janz. 2006. Larval
deformities associated with selenium accumulation in northern pike (Esox lucius) exposed to metal
mining effluent. Environ. Sci. Technol. 40:6506-6512.
Test Organism:
Exposure Route:
Test Duration:
Study Design:
Effects Data:
Northern pike (Esox lucius)
Dietary and waterborne - field exposure
Eggs were collected in the field and incubated in the laboratory. The test was
terminated when the majority of the fry exhibited swim-up and had absorbed the
yolk.
The study area was Key Lake uranium milling operation in north-central
Saskatoon. Spawning northern pike were collected from four sites, one reference
(Davies Creek) and three exposure sites, David Creek near-field (high exposure),
Delta Lake (medium exposure), and David Creek far-field (low exposure). The
exposure sites were located approximately 2, 10 and 15 km downstream of the
effluent discharge. Milt and ova were stripped from ripe fish and eggs were
fertilized in the field. Females were saved for metal analysis and age
determination. Subsamples of ova (prior to fertilization) were collected for metal
analysis.
Although the study sites represent open systems where fish can potentially
migrate among sites, radiotelemetry data from tagged adult pike (Muscatello and
Janz, unpublished data) indicate high site fidelity at the "high" and "medium"
exposure sites (lakes). In contrast, the "low" exposure site likely represents pike
that migrated from further downstream sites that were likely of similar Se
exposures as the reference site.
Eggs were incubated using a two-way ANOVA experimental design using water
collected from reference or exposure sites. So, embryos originating from
reference or exposure site females were incubated in either reference or
appropriate exposure water. In addition, embryos from reference site females
were incubated in water from all four study sites. 50 viable embryos from each
individual female were transferred to each of four replicate incubation chambers.
Cumulative time to 50% eyed, 50% hatch and 50% swim-up were determined.
When the majority of the fry exhibited swim-up and had absorbed the yolk, the
remaining fry were preserved and examined for deformities.
Mean egg diameter and fertilization success did not differ among sites.
Cumulative embryo mortality throughout incubations was not significantly
different among the sites ranging from 45 to 60%. There were no significant
differences in the cumulative time to reach 50% eyed embryos, 50% hatch or
50% swim-up among treatments. Differences in the percent total deformities
between test waters used during embryo incubation exposures were not
significant, so the data were combined for each site (see Table below).
C-24

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Selenium concentrations in eggs and muscle from female northern pike collected from reference
and exposed sites and associated total deformities in embryos


Site
Site ID
Female
fSe] mg/kg dw
Total



Egg
Muscle
deformities %
Davies Creek
Reference
1
3.45
0.86
17
Davies Creek
Reference
2
2.72
1.89
2.5
Davies Creek
Reference
3
3.39
2.56
15.51
Davies Creek
Reference
4
3.72
1.34
7.13
Davies Creek
Reference
5
2.69
1.04
10.41
David Creek (far field)
Low
1
3.39
1.95
20.32
David Creek (far field)
Low
2
4.07
2.04
13.19
David Creek (far field)
Low
3
4.07
1.26
15.33
David Creek (far field)
Low
4
4.07
2.48
18.83
David Creek (far field)
Low
5
3.4
1.26
11.8
Delta Lake
Medium
1
43.19
17
37.8
Delta Lake
Medium
2
24.53
16.52
31.71
Delta Lake
Medium
3
26.14
16.52
26.29
David Creek (near field)
High
1
48.23
47.82
39.5
David Creek (near field)
High
2
N/A*
28.72
N/A*
* female had no eggs
Significant increases in total deformities (edema, skeletal deformities,
craniofacial deformities and fin deformities) were observed in fry originating
from pike collected at the medium exposure site. Determination of an effect level
for the percent total deformities relative to the concentration of selenium in eggs
or in female muscle tissue was not amenable to analysis by TRAP. One
requirement of TRAP is to have a response greater than 50%, which was not
satisfied with the available data.
When data are not amenable to determining an effect level using a software
program, such as TRAP, one way to estimate the effect level is to make a direct
measurement of effect at an exposure or tissue concentration. For example, if
only a control and one exposure concentration, 10 |ag/L, were tested in an acute
toxicity test and there was 100% survival in the control and 35% in the 10 |ag/L,
the effect level would be an EC35 of 10 fxg/L. Such an approach was used to
estimate effect in the Muscatello et al. data. Because no significant differences
were observed in either selenium concentrations in eggs or percent total
deformities between the reference and low exposure site, the data from these 10
sites were combined. Similarly, the egg and muscle selenium and total deformity
data were combined for the 4 medium and high exposure sites. These means,
geometric for the selenium concentrations and arithmetic for the percent total
deformities, are given in the following table.
C-25

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Mean selenium in northern pike egg and muscle and effect values for reference
and exposure sites
Sites
[Se] in eggs,
mg/kg dw
(geometric
mean)
[Se] in muscle,
mg/kg dw
(geometric
mean)
Total deformities, %
(arithmetic mean)
Total deformities,
% (accounting for
reference
deformities and
transformed to
new scale)3
Reference
sites (includes
low exposure)
3.462
1.570
13.20
0
exposure sites
34.00
21.70
33.82
23.76
a The % total del
ormities in the reference and exposed sites were normalized to t
le reference effect
(13.2%) and then transformed to a new scale (100%). i.e, Abbott's formula.
The percent affected becomes 24% or an EC24 and the effect level is 34.00 mg
Se/kg dw in eggs and 21.70 mg Se/kg in muscle.
Chronic Value:	EC24 = 34.00 mg Se/kg dw in eggs. Note: an ECi0 cannot be estimated with the
data.
C-26

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Hamilton, S.J., K.J. Buhl, N.L. Faerber, R.H. Wiedermeyer and F.A. Bullard. 1990. Toxicity of
organic selenium in the diet of chinook salmon. Environ. Toxicol. Chem. 9:347-358.
Test Organism:	Chinook salmon (Oncorhynchus tshawytscha Walbaum; swim-up larvae)
Exposure Route: Dietary only
Control Diet
Oregon moist pellet diet where over half of the salmon meal was replaced with
meal from low-selenium mosquitofish (1.0 mg Se/kg dw) collected from a
reference site.
Selenium Diet # 1
Oregon moist pellet diet where over half of the salmon meal was replaced with
meal from high-selenium mosquitofish (35.4 mg Se/kg dw) collected from the
San Luis Drain, CA, termed SLD diet.
Selenium Diet #2
Oregon moist pellet diet where over half of the salmon meal was replaced with
meal from low-selenium mosquitofish same as in the control diet, but fortified
with seleno-DL-methionine (35.5 mg Se/kg dw), termed SeMet diet.
Dietary Treatments: Each selenium diet was formulated to contain about 36 mg Se/kg dw as the high
exposure treatment. The remaining treatments were achieved by thoroughly
mixing appropriate amounts of high-exposure treatment diet with control diet to
yield the following nominal concentrations (3, 5, 10, and 18 mg Se/kg dw).
Test Duration:	90 days
Study Design:	Each dietary treatment was fed twice each day to swim-up larvae (n=100) in each
of two replicate aquaria that received 1 L of replacement water (a reconstituted
experimental water that simulated in quality a 1:37 dilution of water from the San
Luis Drain, CA minus the trace elements) every 15 minutes (flow-through
design). Mortality was recorded daily. Growth was evaluated at 30-day intervals
by measuring the total lengths and wet weights of two subsets of individual fish
(n=10x2) held in separate 11.5 L growth chambers within each replicate
aquarium. Tissue samples were collected for whole-body selenium
determinations (dw basis) at 30-day intervals throughout the study; 10, 5, and 2
fish were sampled from each duplicate treatment after 30, 60, and 90 days of
exposure, respectively. Concentrations of selenium measured in water were
below the limit of detection (1.5-3.1 |ag/L) in all dietary selenium exposure
concentrations.
C-27

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Effects Data:	The magnitude of reduced growth was most evident in the weight of the fish,
although total length was significantly reduced in fish fed high Se-laden diets as
well. The effect of increasing dietary selenium on mean larval weight was similar
in both the SLD and seleno-methionine diets.
Effect of San Luis Drain Diet on Growth and Survival of Chinook Salmon Larvae after 60 Days
Se in diet, mg/kg dw
Se in chinook salmon,
mg/kg dw
Mean larval weight, g
Survival, %
1
0.9
3.35
99
3.2
3.3
2.68
97.3
5.3
4.5
2.76
93
9.6
8.4
2.8
95
18.2
13.3
2.62
92.4
35.4
29.4
1.4
89

Effect of Seleno-methionine Diet on Growth and Survival of Chinook Salmon Larvae after 60
Days
Se in diet, mg/kg dw
Se in chinook salmon,
mg/kg dw
Mean larval weight, g
Survival, %
1
0.9
3.35
99
3.2
2
3.08
100
5.3
3.1
3.22
95
9.6
5.3
3.07
94.1
18.2
10.4
2.61
92.4
35.4
23.4
1.25
62.5
C-28

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Chronic Value:	Due to unacceptable control mortality of swim-up larvae in control treatments
after 90 days (33.3 percent - SLD diet; 27.5 percent - SeMet diet), chronic values
had to be determined from respective values reported after 60 days (tables
above).
Analysis of the elemental composition of the SLD diet indicated that B, Cr, Fe,
Mg, Ni and Sr were slightly elevated compared to the control and SeMet diets.
No additional analyses were performed to determine the presence of other
possible contaminants, i.e., pesticides.
Diet
type
EC2o values
ECio values
Survival
(after 60 d of
exposure)
Growth
(after 60 d of exposure)
Growth
(after 60 d of exposure)
Tissue Se
(mg/kg dw)
Whole body Tissue Se
(mg/kg dw)
Whole body Tissue Se
(mg/kg dw)
SLD
NAa
15.73
11.14
SeMet
NAa
10.47
7.355
a The EC2o and ECi0 values for survival of swim-up larvae versus levels of selenium for the SLD and
SeMet dietary exposure could not be estimated using non-linear regression.
C-29

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Hamilton et al (1990) Chinook Salmon fed SLD Diet
Logistic Equation, Three Parameter Model, Se concentrations logio transformed
O)
'<5
4.0
3.5
3.0
2.5
= 2.0
CO
>
CO
CO
CD
1.5
1.0
.5

U
0
.2
.4
O
CO
CO



Log(Se in Chinook Salmon mg/kg

Guess
FinalEst
SE
95% LCL 95%UCL
LogX50
1.453
1.453
7.30E-02
1.2206 1.6854
StDev
1.353
1.353
6.67E-01
-7.71E-01 3.4769
YO
2.968
2.968
1.89E-01
2.3651 3.5709
% Effect
Xp Est
95% LCL
95% UCL

50
28.379
16.62
48.458

20
15.734
5.7003
43.431

10
11.143
2.4771
50.127

5
8.1085
1.1213
58.637


DF
SS
MS
F P
Total
5
2.0749
0.41498

Model
2
1.8202
0.91009
10.719 0.95699
Error
3
0.2547
8.49E-02

1.2
1.4
1.6
C-30

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Chinook salmon SeMet diet (Hamilton et al. 1990)
_
>
to
1.6
1.4
.8
.8
1.0
1.2
0
.2
.4
Log([Se]whole body, mg/kg dw)
Parameter Summary (Logistic Equation Regression Analysis)
Parameter
Guess
FinalEst
StdError
95%LCL
95%UCL
LogXSO
1.3148
1.2823
0.0242
1.2053
1.3593
S
0 6971
1.3214
0.1826
0.7404
1.9025
Y0
3.217
3.239
0.067
3.027
3.452

Effect Concentration Summary

% Effect
Xp Est
95%LCL
95%UCL
50.0
19.156
16.045
22.870
20.0
10.472
7.516
14.591
10.0
7.355
4.595
11.775
5 0
5.312
2.899
9.733
06/19/2009 13:52	MED Toxic Response Analysis Model, Version 1 03
C-31

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Hilton, J.W. and P.V. Hodson. 1983. Effect of increased dietary carbohydrate on selenium metabolism
and toxicity in rainbow trout (Salmo gairdneri). J. Nutr. 113:1241-1248.
Test Organism:	Rainbow trout (Oncorhynchus mykiss; juvenile; approx. 0.6 g each)
Exposure Route: Dietary only
Low carbohydrate diet (LCD)
This diet contained capelin oil at 11 percent of the diet with cellulose as the filler.
High carbohydrate diet (HCD)
This diet contained cerelose at 25 percent of the diet with cellulose as the filler.
For both diets, the selenium was supplemented as sodium selenite which was
mixed with cellulose and then added to the diet as a selenium premix.
Test Treatments: The two diets were supplemented with selenium (as sodium selenite) at the rate
of 0, 5, or 10 mg/kg dw to make up the six different dietary selenium treatments
(n = 3 low carbohydrate diet; n= 3 high carbohydrate diet). The six diets were fed
to duplicate groups of 100 fish. The trout were fed to satiation 3-6 times per day.
Measured concentrations of selenium in the low carbohydrate diet were: 0.6
(control), 6.6, and 11.4 mg/kg dw, and the measured concentrations of selenium
in the high carbohydrate diet were: 0.7 (control), 6.6, and 11.8 mg/kg dw. The
tanks received a continuous flow of water with a flow rate of 3-4 liters per
minute.
Test Duration:	16 weeks
Study Design:	Body weights, feed: gain ratios, and total mortalities were determined after each
28-day interval. After 16 weeks, approximately 20 fish were randomly removed
from each tank, weighed, and blood was collected for hemoglobin, hematocrit,
and plasma glucose, protein, and calcium determination. The livers and kidneys
were then dissected. The livers were assayed for glycogen content, and samples
of both liver and kidney were assayed for selenium content. Additional
subsamples of fish were sacrificed and assayed for selenium content and for ash,
crude protein, and moisture content (n=6 per treatment). Finally, 30 fish were
killed, their livers and kidneys dissected, and analyzed for Ca, Cu, Fe, Mg, P, and
Zn content.
Effects Data:	The only overt sign of selenium toxicity was food avoidance observed in trout
fed the highest selenium content in both low and high carbohydrate diets, which
led to significantly reduced body weight after 16 weeks. There were no
significant differences detected between treatment groups in hematological
parameters. Kidney, liver, and carcass selenium levels increased with increasing
selenium content of the diet, however, only the liver selenium concentrations
were significantly affected by dietary selenium level, dietary carbohydrate level,
and the interaction between the two treatments. Mineral analysis of the kidney
showed significantly higher levels of calcium and phosphorous in trout reared on
the two highest levels of dietary selenium. Concentrations of copper in the liver
increased significantly with increasing dietary selenium levels and decreasing
dietary carbohydrate levels.
C-32

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Effect of Selenium in Low carbohydrate Diet to Rainbow Trout
Se in diet, mg/kg dw
Se in trout liver, mg/kg dw
Trout weight, kg/100 fish
0.6
0.8
3.3
6.6
38.3
3.3
11.4
49.3
1.8

Effect of Selenium in High carbohydrate Diet to Rainbow Trout
Se in diet, mg/kg dw
Se in trout liver, mg/kg dw
Trout weight, kg/100 fish
0.7
0.6
2.7
6.6
21.0
2.3
11.8
71.7
1.4
Chronic Value:	The following table lists the NOAEC, LOAEC and MATC for both diets in liver
tissue. EC values could not be determined for this study. Data did not meet
minimum requirements for analysis.
Diet
NOAEC, mg Se/kg dw
liver
LOAEC, mg Se/kg dw
liver
MATC, mg Se/kg dw
liver
Low carb
38.3
49.3
43.5
high carb
21.0
71.7
38.8
C-33

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Hicks, B.D., J.W. Hilton, and H.W. Ferguson. 1984. Influence of dietary selenium on the occurrence of
nephrocalcinosis in the rainbow trout, Salmo gairdneri Richardson. J. Fish Diseases. 7:379-389.
(Note: These data are the exact same as reported for the low carbohydrate diet in Hilton and Hodson
1983, with the addition of prevalence of nephrocalcinosis occurring in trout after 16 to 20 weeks of
consuming the contaminated test diets).
Test Organism:
Exposure Route:
Test Treatments:
Test Duration:
Study Design:
Effects Data:
Rainbow trout (Oncorhynchus mykiss; juvenile; approx. 0.6 g each)
Dietary only
This diet contained capelin oil at 11 percent of the diet with cellulose as the filler.
The selenium was supplemented as sodium selenite which was mixed with
cellulose and then added to the diet as a selenium premix.
The test diet was supplemented with selenium (as sodium selenite) at the rate of
0, 5, or 10 mg/kg dw to make up the three different dietary selenium treatments.
The three diets were fed to duplicate groups of 100 fish. The trout were fed to
satiation 3-6 times per day. Measured concentrations of selenium in the low
carbohydrate diet were: 0.6 (control), 6.6, and 11.4 mg/kg dw. The tanks received
a continuous flow of water with a flow rate of 3-4 liters per minute.
16 to 20 weeks
See Hilton and Hodson (1983). After 20 weeks on the test diets, ten fish were
randomly removed from each treatment. Tissues for histopathological
examination included the stomach, intestine and pyloric ceca (including
pancreas), spleen, liver, heart, kidney, skin, muscle, and gills.
Only effects of selenium on kidney tissue are included in the article. The kidneys
of the 10 trout fed the highest selenium content in the diet exhibited normal
appearance. Five of these trout exhibited precipitation of calcium in the tubules
with some epithelial necrosis, but no loss of epithelial continuity. Extensive
mineralized deposition of Ca within the tubules, tubular dilation and necrosis of
tubular epithelium, ulceration of tubules, and intestinal Ca mineralization was
observed in four of the ten fish.
Chronic Value:	Same as for growth of rainbow trout reported by Hilton and Hodson (1983). The
MATC estimated for growth of rainbow trout relative to final concentration of
selenium in liver tissue of trout reared on the low carbohydrate diet is the GM of
38.3 (NOAEC) and 49.3 (LOAEC) mg/kg dw, or 43.45 mg/kg dw.
EC values could not be determined for this study. Data did not meet minimum
requirements for analysis.
C-34

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Hilton, J.W., P.V. Hodson, and S.J. Slinger. 1980. The requirements and toxicity of selenium in
rainbow trout (Salmo gairdneri). J. Nutr. 110:2527-2535.
Test Organism:	Rainbow trout (Oncorhynchus mykiss; juvenile; approx. 1.28 g each)
Exposure Route: Dietary only
A casien-torula yeast diet was formulated to contain geometrically increasing
levels of selenium from 0 to 15 mg/kg dw. The selenium was supplemented as
sodium selenite which was mixed with cellulose and then added to the diet as a
selenium premix.
Test Duration:	20 weeks
Study Design:	Six test diets were fed to triplicate groups of 75 fish. The trout were fed to
satiation 3-4 times per day, 6 days per week, with one feeding on the seventh
day. Measured concentrations of selenium in the diet were: 0.07 (control), 0.15,
0.38, 1.25, 3.67, and 13.06 mg/kg dw. The tanks received a continuous flow of
dechlorinated tap water from the City of Burlington, Ontario municipal water
supply. The waterborne selenium content of this water was 0.4jli g/L. During the
experiment, the fish were weighed every 2 weeks with the feeding level adjusted
accordingly. Mortalities were noted daily and the feed consumption for each
treatment was recorded weekly. After 4 and 16 weeks, three to six fish were
randomly removed from each tank, sacrificed, and their livers and kidneys
removed and weighed. An additional three to six fish were then obtained from
each treatment, killed, and prepared for tissue analysis. Organs and carcasses
were freeze-dried for determination of selenium concentration. After 16 weeks,
three more fish were removed. Kidney, liver, spleen and dorsal muscle tissue was
dissected for examination of histopathology. At the end of 8 and 16 weeks, four
to five fish were removed, sacrificed, and a blood sample was taken for
hematological measurements (hematocrit, red blood cell count, and blood iron
concentration). After 20 weeks, three to four more fish were removed, sacrificed,
and a blood sample was taken for measurement of glutathione peroxidase
activity.
Effects Data:	There were no significant differences detected between treatment groups in
histopathology, hematology, or plasma glutathione peroxidase activity. Trout
raised on the highest dietary level of selenium (13.06 mg/kg dw) had a
significantly lower body weight and a higher number of mortalities (10.7;
expressed as number per 10,000 fish days) than trout from the other treatments
levels after 20 weeks of exposure.
C-35

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Effects on Juvenile Rainbow Trout
Se in diet, mg/kg dw
Se in Liver, mg/kg dw
Weight, g/fish
Mortality*
0.07
0.6
3.2
0
0.15
0.95
3.5
0.38
2.4
3.7
0.6
1.25
11
4.1
0.6
3.67
40a
4.1
13.06
100D
1.4
10.7
expressed as number per 10,000 fish-days
NOAEC
LOAEC
Chronic Value:	NOAEC = 40 mg Se/kg dw
LOAEC =100 mg Se/kg dw
MATC = 63.25 mg Se/kg dw
C-36

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Holm, J. 2002. Sublethal effects of selenium on rainbow trout (Oncorhynchus mykiss) and brook trout
(Salvelinus fontinalis). Masters Thesis. Department of Zoology, University of Manitoba, Winnipeg, MB.
Holm, J., V.P. Palace, K. Wautier, R.E. Evans, C.L. Baron, C. Podemski, P. Siwik and G. Sterling.
2003. An assessment of the development and survival of rainbow trout (Oncorhynchus mykiss) and brook
trout (Salve linus fontinalis) exposed to elevated selenium in an area of active coal mining. Proceedings of
the 26th Annual Larval Fish Conference 2003, Bergen, Norway. ISBN 82-7461-059-B.
Holm, J., V.P. Palace, P. Siwik, G. Sterling, R. Evans, C. Baron, J. Werner, and K. Wautier. 2005.
Developmental effects of bioaccumulated selenium in eggs and larvae of two salmonid species. Environ.
Toxicol. Chem. 24: 2373-2381.
Test Organism:
Exposure Route:
Study Design:
Effects Data
Rainbow trout (Oncorhynchus mykiss; spawning adults) and brook trout
(Salvelinus fontinalis; spawning adults)
Dietary and waterborne - field exposure
Total selenium concentrations measured at the high selenium site ranged from 6
to 32 |ag/L. Selenium was not measured at the reference streams; selenium
concentrations at reference locations in the area ranged from <0.5 to 2.2 |ag/L.
Spawning fish were collected at low selenium or reference streams (Deerlick
Creek, Wampus Creek and Cold Creek), a slightly elevated selenium stream
(Gregg Creek), and an elevated selenium stream (Luscar Creek) in the
Northeastern slopes region of Alberta, Canada. An active coal mine is the source
of selenium in the elevated streams. Eggs and milt from the spawning trout were
expressed by light pressure from abdomen. Individual clutches of eggs were
fertilized from a composite volume of milt derived from 3-5 males. Fertilized
eggs from individual females were reared to swim-up stage and examined for a
number of parameters including percent fertilization, mortality, edema, and
deformities (craniofacial, finfold, and spinal malformations). Similar studies
were conducted in 2000, 2001 and 2002. One notable difference is that the
embryos were incubated at 8C in 2000 and at 5C in 2001. The authors noted
that 5C is a better representation of the actual stream temperature during embryo
development.
Other than selenium, there were no significant differences in the concentrations
of other elements (Al, As, Sb, Ba, Be, Ni, B, Cd, Ca, Cr, Co, Cu, Fe, Pb, Li, Mg,
Mn, Hg, Mo, Ag, Sr, Tl, Th, Sn, Ti, U, V, Zn) in trout eggs between the low level
and elevated selenium streams. There are two ways to approach determination of
effects due to selenium in this study and both are presented here. The first
approach determines effects based on a comparison of average conditions
between streams (between streams approach). For example, if there is a
significant difference between the average frequency of deformities in a
contaminated stream and reference stream, the effect level for the between
streams approach would be the average concentration of selenium in the tissue
from the contaminated stream. The second approach evaluates individual
response variables (e.g., edema, deformities) against the individual selenium
tissue concentrations for the combined contaminated and reference stream data
set with each year (within streams approach). This approach, which results in an
C-37

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EC estimate (e.g., ECi0) if the data meet the model assumptions, is explained
below.
Between streams approach: For each sampling location (stream), data for the
three years (Tables 1 and 2) were combined in the between streams analysis of
variance (ANOVA). For rainbow trout embryos, there were no significant
differences in fertilization, time to hatch and mortality between the streams with
elevated selenium and the reference streams. ANOVA indicated significant
differences in the frequency of embryonic effects between streams (Table 3). The
analysis did not prove useful; however, due to a higher occurrence of effects in
some of the reference streams relative to the exposed streams (Tables 3 and 4).
The between streams analysis, therefore, was not used to determine effect
concentrations for rainbow trout.
ANOVA of brook trout data indicated the only significant difference in
embryonic abnormalities among sites was craniofacial deformities (Tables 5 and
6). Significant differences were also found for fertilization and larval weight. The
highest average percent fertilization was observed at the site with the greatest
concentration of selenium in eggs, which indicates that the differences in
fertilization among sites were not caused by variation in selenium concentrations.
Because the percent of embryos with craniofacial deformities in Luscar Creek
was 7.9% (2.1% in Cold Creek), it was not considered biologically meaningful.
Likewise the significantly lower larval weights at the exposed sites was not large
(16% lower than Cold Creek larvae) and again coupled with the low occurrence
of abnormalities by the brook trout, a signature of selenium effects, the lower
larval weights were not considered biologically meaningful.
Within streams approach: As with the between streams analysis, data were
combined for the three years of study in the within streams analysis (Tables 1 and
2). Craniofacial deformities, skeletal deformities and edema in rainbow trout
embryo, as a function of selenium in egg ww, were fitted to a curve using a
weighted regression and threshold sigmoidal equation from which ECi0 values
were calculated (see Figures 1, 2 and 3). EC estimates for finfold deformities,
length and weight of rainbow trout embryos could not be made because of
inadequate dose-response. The brook trout data were not suitable for fitting
logistic curves (Figure 5).
C-38

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Table 1. Rainbow trout embryo-larval parameters collected from a high Se site (Luscar Creek), an
intermediate Se site (Gregg River), and reference sites (Deerlick Creek and Wampus Creek) in
northeastern Alberta over three consecutive years.
Year Site Female # Se in eggs, %craniofacial %skeletal %finfold %edema
	mg/kg ww deformities deformities deformities	
2000
Luscar
11
6.84
7.18
13.26
1.66
4.97
2000
Luscar
12
6.66
1.48
4.43
0.74
1.85
2000
Luscar
14
11.6
14.43
23.71
7.22
85.57
2000
Deerlick
16
1.78
0.63
1.9
0.63
0.63
2000
Deerlick
17
1.39
0
0
0
0
2000
Deerlick
18
1.00
0
0.86
0
0
2000
Deerlick
15
5.01
0
0
0
0
2001
Luscar
1
5.39
7.35
6.76
3.53
2.94
2001
Luscar
3
8.39
6.29
4.97
2.98
6.95
2001
Luscar
4
6.48
22.22
22.22
33.33
26.67
2001
Luscar
8
4.47
12
9.33
2.67
10.67
2001
Luscar
14
10.4
34.55
44.85
4.24
43.64
2001
Luscar
32
5.64
8.24
5.97
3.13
9.09
2001
Luscar
33
3.88
5.26
6.58
9.21
3.95
2001
Luscar
39
5.14
1.91
3.18
0
1.27
2001
Luscar
40
3.36
11.62
7.05
5.39
6.64
2001
Luscar
41
11.7
37.67
83.41
3.59
87
2001
Deerlick
8
3.68
9.55
5.45
1.36
5.45
2001
Deerlick
9
3.08
5.39
4.98
0.41
2.07
2001
Deerlick
10
1.62
7.89
7.89
5.26
10.53
2001
Deerlick
16
2.62
24.24
48.48
3.03
12.12
2001
Deerlick
17
2.79
14.13
15.22
4.35
20.65
2001
Deerlick
21
1.96
13.27
35.71
7.14
25.51
2001
Deerlick
22
3.13
1.09
2.17
0
1.09
2001
Deerlick
23
3.03
9.65
14.04
3.51
7.89
2001
Deerlick
25
3.32
9.25
13.29
7.51
8.09
2001
Deerlick
39
2.43
11.89
9.09
7.69
14.69
2001
Gregg
2
4.57
11.97
7.75
15.49
7.04
2001
Gregg
3
4.49
5.58
9.3
2.33
4.65
2001
Gregg
5
4.05
4.95
5.45
2.48
5.94
2001
Gregg
9
5.09
20
13.85
15.38
16.15
2001
Gregg
18
5.97
16.13
19.35
41.94
35.48
2001
Wampus
9
2.66
16.07
0
1.79
7.14
2001
Wampus
13
2.04
7.84
9.8
1.31
7.84
2002
Luscar
3
5.4
60.47
27.9
93
14
2002
Luscar
8
18.3
94.12
23.5
4.4
97.1
2002
Luscar
10
22
100
64.3
3.6
100
2002
Luscar
12
15.7
82.35
47.1
66.7
52.9
C-39

-------
Year
Site
Female #
Se in eggs,
mg/kg ww
%craniofacial
deformities
%skeletal %finfold
deformities deformities
%edem
2002
Luscar
22
20.5
100
42.1
2.1
100
2002
Luscar
23
6.3
5.59
6.6
1.6
2.7
2002
Luscar
24
26.8
100
100
0
100
2002
Luscar
26
6.5
1.72
1.7
4.3
0.9
2002
Deerlick
10
5.9
5.65
7.26
7.26
3.23
2002
Deerlick
18
7.8
10.77
1.54
9.23
3.08
2002
Deerlick
21
5
6.9
6.9
20.69
1.72
2002
Deerlick
24
4.3
2.88
2.88
21.58
0.72
2002
Deerlick
25
4.4
5.3
5.3
6.82
3.03
2002
Deerlick
26
6.6
2.95
1.85
1.11
1.85
2002
Gregg
1
5.8
4.76
3.81
3.81
3.81
2002
Wampus
1
3
18.84
14.49
72.46
11.59
2002
Wampus
2
4
0
0
100
100
2002
Wampus
3
4.6
4.1
3.28
7.58
0.61
2002
Wampus
4
4.7
25
20
70
12.5
2002
Luscar
28
7
19.23
0
76.9
0
C-40

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Table 2. Brook trout embryo-larval parameters collected from a high Se site (Luscar Creek), an
intermediate Se site (Gregg River), and reference site (Cold Creek) in northeastern Alberta over
three consecutive years.
Year Location Female # Se in egg, mg/kg %craniofaci %skeletal %finfold %edema
2000
Luscar
1
4.78
15.38
0
0
15.38
2000
Luscar
2
4.83
38.06
1.49
3.73
1.49
2000
Luscar
3
5.98
7.39
3.03
0.34
0.5
2000
Luscar
5
3.86
25
5.7
8.77
4.82
2000
Luscar
12
6.06
16.77
1.83
0.7
0
2000
Luscar
13
5.8
4.06
1.42
0.2
0
2000
Luscar
14
5.17
4.13
0.49
0.36
0.12
2000
Luscar
15
9.92
16.22
0.54
0.54
0
2000
Luscar
16
5.03
5.61
0
0.27
0.27
2000
Luscar
17
6.01
9.44
5.83
0.83
1.11
2000
Luscar
18
12.7
14.34
0.72
0
0.36
2000
Cold
21
1.15
3.26
1.48
0.89
0
2000
Cold
22
1.83
4.83
1.38
1.38
0.69
2000
Cold
24
0.97
1.67
0
0.72
0
2000
Cold
25
No data
3.31
1.1
1.66
1.1
2000
Cold
26
0.59
3.45
4.83
6.9
0.69
2000
Cold
33
1.35
6.15
0
1.54
0
2000
Cold
34
2.18
6.45
0
0.81
0
2001
Cold
6
1.79
0
0
0
0
2001
Cold
7
1.36
1.61
0.69
0.46
1.38
2001
Cold
8
0.94
1.36
0
0.27
0.54
2001
Cold
21
1.07
0.43
0
0
0
2001
Cold
51
1.09
0
2.13
0
6.38
2001
Luscar
3
8.4
0
0.93
0
0.46
2001
Luscar
7
7.26
1.35
1.62
0.81
0.27
2001
Luscar
17
14.6
2.22
0.63
0.32
0
2001
Luscar
19
9.79
7.55
2.11
2.42
0.3
2001
Luscar
59
5.8
2.28
0.46
0.91
0.46
2001
Luscar
60
9.03
3.16
0
1.05
1.05
2001
Luscar
61
7.29
0
0
9.09
0
2001
Luscar
64
7.08
1.54
2.19
0
0
2001
Luscar
76
7.1
36.71
13.29
19.65
1.16
2001
Luscar
82
6.06
1.11
0.22
0.88
0.44
2001
Luscar
83
5.82
6
2
5.6
0.8
2001
Gregg
3
7.08
6.32
1.58
20.53
1.58
2001
Gregg
22
7.95
0
0
1.08
0
2001
Gregg
23
9.23
0.5
0.5
2.51
0
2001
Gregg
25
6.46
0.56
0
0.56
0
C-41

-------
Year Location Female # Se in egg, mg/kg %craniofaci %skeletal %finfold %edema
2001
Gregg
2001
Gregg
2001
Gregg
2001
Gregg
2002
Luscar
2002
Luscar
2002
Luscar
2002
Luscar
2002
Luscar
2002
Luscar
2002
Luscar
2002
Luscar
2002
Gregg
2002
Gregg
2002
Gregg
2002
Cold
2002
Cold
2002
Cold
2002
Cold
2002
Cold
2002
Cold
2002
Cold
2002
Cold
2002
Cold
2002
Cold
31
7.35
32
4.91
33
7.02
34
5.01
17
6.28
23
5.27
26
6.36
38
18.9
42
4.95
44
6.47
54
7.96
56
18.8
25
6.27
37
4.58
39
6.67
32
0.42
26
0.89
2
0.94
5
1
29
1.02
23
1.2
48
1.25
42
1.6
22
1.74
51
2.11
0.51
1.7
7.21
0.48
1.88
1.88
0
0.37
1.7
12.74
7.34
0.46
1.81
0.52
0.9
0.54
2.79
0.44
0
0.25
0.33
0.33
3.99
0.75
1.23
1.23
2.99
0
3.57
1.19
0
0.6
0
0
0.96
0.32
0.25
0.5
0.72
1.09
0.35
0.35
9.52
4.76
0
0
0
0
2.17
2.17
0.17
0
3.37
0.48
4.38
0
0
0
0.85
0.21
0
0.46
0.26
0.26
0
0.18
0.15
0.15
0
0
0
0
0.5
0.75
0
0
0
0
1.19
1.19
0
0
0
0.29
0
0
0.25
0
0.36
0.72
0.35
0.35
2.38
0
0
0
1.09
1.09
0
2.17
C-42

-------
Table 3. Results of ANOVA comparing rainbow trout endpoints among sites
% fertilization

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
77.60
25.8653
0.06336703
0.978935
Residuals
51
20817.33
408.1829


% mortality

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
3751.51
1250.504
1.848008
0.1502207
Residuals
51
34510.50
676.676


% craniofacia
deformities

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
8093.97
2697.989
4.430272
0.007732133
Residuals
50
30449.48
608.990


% skeletal deformities

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
3279.30
1093.101
2.773923
0.05094422
Residuals
50
19703.16
394.063


% finfold deformities

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
6273.17
2091.056
3.888612
0.01417887
Residuals
50
26886.93
537.739


% edema

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
8902.51
2967.502
3.449597
0.0233558
Residuals
50
43012.30
860.246


C-43

-------
Table 3. Results of ANOVA comparing rainbow trout endpoints among sites (continued)
Fry length 					

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
5.0847
1.694896
0.5694271
0.6377436
Residuals
50
148.8246
2.976493


Fry weight

Df
Sum of Sq
Mean Sq
F Value
Pr(F)
Site
3
1721.104
573.7012
3.563888
0.02080915
Residuals
48
7726.859
160.9762


Table 4. Rainbow trout means (standard deviation) for measurements made in eggs, embryos and
larvae spawned from fish collected at exposed sites (Luscar and Gregg Creeks) and reference sites
(Deerlick and Wampus Creeks).	
Parameter
Site
Luscar Cr.
Gregg Cr.
Deerlick Cr.
Wampus Cr.
egg Se, mg/kg ww
9.93 (6.77)
6.52 (4.11)
3.49 (1.90)
3.5 (1.09)
fertilization, %
77.8 (20.3)
81.2 (12.7)
77.5 (20.9)
77.5 (24.1)
mortality, %
35.0 (29.5)
34.2 (32.5)
18.1 (14.6)
37.3 (34.5)
craniofacial, %
33.3 (37.2)
10.6 (6.5)
7.1(6.1)
12.0 (9.6)
skeletal, %
25.0(27.9)
9.9 (5.8)
9.2 (12.3)
7.9 (8.2)
finfold, %
15.0(27.1)
13.6 (15.2)
5.4 (6.2)
42.2 (43.7)
edema, %
34.5 (40.3)
12.2 (12.3)
6.1 (7.3)
23.3 (37.8)
larval length, mm
18.5 (2.0)
19.4 (1.6)
19.0 (1.5)
19.2 (0.9)
larval weight, mg
53.3 (16.3)
44.6 (10.4)
41.2 (9.3)
40.6 (8.4)
C-44

-------
Table 5. Brook trout means (standard deviation) for measurements made in eggs, embryos and
larva spawned from fish collected at exposed sites (Luscar and Gregg Creeks) and
Parameter
Site
Luscar Cr.
Gregg Cr.
Cold Cr.
egg Se, mg/kg ww
7.78 (3.80)
6.59 (1.39)
1.26 (0.47)
fertilization, %
92.8 (7.2)
78.4(18.2)
89.1 (19.6)
mortality, %
6.5 (8.9)
2.9 (2.3)
6.9(12.1)
craniofacial, %
7.9 (10.1)
2.3 (2.5)
2.1 (2.6)
skeletal, %
2.0 (3.3)
0.8 (0.7)
1.0(1.4)
finfold, %
1.9 (4.1)
3.1 (6.0)
0.9(1.5)
edema, %
1.0 (2.9)
0.3 (0.6)
0.7(1.4)
larval length, mm
17.4 (1.1)
17.9(0.9)
18.5 (1.2)
larval weight, mg
31.7 (8.6)
31.3 (5.4)
37.8 (7.2)
% fertilization






df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
1683.3
841.67
3.9128
0.0253
Residuals
60
12906.4
215.11








% mortality






df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
131.4
65.72
0.7257
0.4882
Residuals
60
5433.6
90.56








% craniofacial deformities




df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
519.1
259.54
4.9427
0.0103
Residuals
60
3150.6
52.51


C-45

-------
Table 6. Results of ANOVA comparing brook trout endpoints among sites (continued)
% skeletal deformities





df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
19.2
9.58
1.5631
0.2179
Residuals
60
367.6
6.13








% finfold deformities





df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
37.5
18.74
1.2562
0.2921
Residuals
60
895.1
14.92








% edema






df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
4.6
2.32
0.4966
0.6110
Residuals
60
280.6
4.68








Fry length






df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
16.1
8.04
6.5265
0.0027
Residuals
60
73.9
1.23








Fry weight






df
Sum of Sq
Mean Sq
F Value
Pr(F)
site
2
546.2
273.10
4.6644
0.0131
Residuals
60
3512.9
58.55


C-46

-------
100'
80 
CO
0)
'-I'
'E
o
4
CD
Q
CC 60 -
o
cc

O
o
o
40 -
20 -
 

 
-  
. V v.
10
20
mg Se/kg egg
Figure 1. Rainbow trout percent normal (100 - % craniofacial deformities) as a function of the
logarithm of selenium concentration in eggs wet weight (Exposure Variable). TRAP weighted
regression analysis using a threshold sigmoid equation. The background value was estimated to be 90.2%,
the slope 4.8%, and the ECio 10.2 mg Se/kg egg ww.
100 i
(/)
U
~ 80
60

-------
1	2	5	10	20
mg Se/kg egg
Figure 3. Rainbow trout percent normal (100 - % edema) as a function of the logarithm of selenium
concentration in eggs wet weight (Exposure Variable). TRAP weighted regression analysis using a
threshold sigmoidal equation excluding the one outlier with 100% edema at 4 mg/kg. The background
value was estimated to be 92.8%, the slope 4.6%, and the ECi0 9.5 mg Se/kg egg ww.
The previous draft used a TRAP logistic regression (Figure 4). A weighted regression using a threshold
sigmoidal equation (Figures 1-3) is a better application of these data.
120
in

-------
Figure 5. Plot of percent abnormal for craniofacial, skeletal and finfold deformities and edema
against selenium concentration in brook trout eggs ww, 2000 and 2001 data.
 ~
vtro
.'Vm *
w
.a;
HI
E
L_
>2
a;
CD
-i
a;
w
14 1

12 -

10 -

8 -

6 -
V
4 -

2 -
o

0 -
Vom
vo

8 10 12 14 16
8 10 12 14 16

8 10 12 14 16
Egg Se concentration
(ing/g, wet weight)
7 i
6 -
5 -
4 
CD
E 3H
0
B 2 -
1 -
0 -

Luscar 2001
o
Cold 2001
~
Gregg 2001
V
Cold 2000

Luscar 2000
~
 
^ . T. .  #
woan tWw ~
i	1
0 2
8 10 12 14 16
Egg Se Concentration
(ing/g, wet weight)
C-49

-------
The effect levels determined using the within streams approach resulted in values based on ww in eggs.
The primary tissue for which the reproductive effect levels were based, eggs, was converted from ww to
dw using the average percent moisture of 61.2% for rainbow trout eggs reported by Seilor and Skorupa
(2001).
Chronic Values: Brook trout: Between streams approach
No effects at ECi0 level at 7.78 mg Se/kg eggs ww or 20.05 mg Se/kg eggs dw;
egg. Chronic value is >20.05 mg Se/kg eggs dw. Table 3 data, converted to dry
weight, suggest no effects at least up to 25-35 mg Se/kg eggs dw.
Rainbow trout: Within streams approach
ECio value (edema) at 9.5 mg Se/kg egg ww or 24.5 mg Se/kg egg dw. Chronic
value is 24.5 mg Se/kg eggs dw.
C-50

-------
Kennedy, C.J., L.E. McDonald, R. Loveridge, M.M. Strosher. 2000. The effect of bioaccumulated
selenium on mortalities and deformities in the eggs, larvae, and fry of a wild population of cutthroat trout
(Oncorhynchus clarkii lewisi). Arch. Environ. Contam. Toxicol. 39:46-52.
Cutthroat trout (Oncorhynchus clarkii lewisi; spawning adults, 3-6 years)
Dietary and waterborne - field exposure
Total selenium concentrations measured at the time the eggs were taken were
<0.1 |jg/L from the reference site and 13.3 to 14.5 |a,g/L at the exposed site.
At reference and exposed site (Fording River, BC, Canada which receives
drainage from open-pit coal mining), eggs were stripped from females (n=20
from reference site; n=17 from exposed site) and fertilized from milt from one
male collected at each site. Fertilized eggs were reared in well water and
examined for time to hatch, deformities (craniofacial, finfold, skeletal and yolk
sac malformations), and mortalities. Inspection of deformities in eggs was
performed using 40X magnification.
No significant correlations between the selenium concentrations in the eggs from
either site and: hatching time (reference, 25.5-26.5 days; exposed, 22-25.5 days);
percent deformities preponding (reference, 0-2.4%; exposed, 0-0.34%); percent
deformities after ponding (reference, 0-0.26%; exposed, 0-0.09%); percent
mortalities preponding (reference, 1.5-70.3%; exposed, 1-100%); percent
mortalities after ponding (reference, 0.3-4.3%; exposed, 1.5-43.7%); total percent
mortalities (reference, 2.8-55.8%; exposed, 3.7-100%). The average selenium
residues in tissues were as follows:
Site
Adult fish liver, mg Se/kg
dw
Adult fish muscle, mg Se/kg
dw
eggs, mg Se/kg dw
Reference
8.2; Range: 3.4-14.6
2.4; 1.4-3.8
4.6
Exposed
36.6; Range: 18.3-114
12.5; Range: 6.7-41
21.2
Chronic Value:	>21.2 mg Se/kg dw in eggs
>12.5 mg Se/kg dw in muscle
C-51
Test Organism:
Exposure Route:
Study Design:
Effects Data

-------
Hardy, R.W. 2005. Effects of dietary selenium on cutthroat trout (Oncorhynchus clarkii) growth and
reproductive performance. Report for Montgomery Watson Harza. December 14, 2005.
Test Organism:	Cutthroat trout (Oncorhynchus clarkii, 0.9 g)
Exposure Route: Dietary only
Six experimental dietary treatments were produced by cold extrusion. The
formulation of the diet was designed to be similar to commercial trout diets and
had a proximate composition of 45% protein and 16% lipid. Seleno-methionine
diluted in distilled water (100 j^ig/L) was added in appropriate volumes to each
batch of feed to facilitate pelleting. Measured dietary selenium concentrations
were 1.2 (control), 3.8, 6.4, 9.0, 11.5, and 12 mg Se/kg dw. Fry were fed initially
at a rate of 10 times per day 6 days each week to apparent satiation. Feeding
frequency decreased as fish grew.
Test Duration:	124 weeks (865 days, 2.5 yrs)
Study Design:	Groups of 50 fish were placed into triplicate tanks (145 L) receiving 4-15 L/min
of hatchery water at 14.5EC and fed one of the six experimental diets. The fish in
each tank were bulk-weighed and counted every 14 days for the first 12 weeks of
the experiment, and then every 4 weeks until 48 weeks. Samples of fish for
whole-body selenium analysis were taken at each sampling date for the first 12
weeks followed by every 3 months thereafter. After six months of feeding, the
fish were transferred to 575 L tanks and the number of replicate tanks per dietary
treatment was reduced to two. After 80 weeks of feeding, the fish were
transferred to 1050 L outdoor tanks each supplied with 70 L/min of constant
temperature (14.5C) spring (hatchery) water. After 2.5 years of the feeding trial,
fish were spawned and whole body selenium level, egg selenium level, % eyed
eggs, % hatched eggs, and % deformed larvae were examined.
Effects Data:	No signs of toxicity (reduced growth or survival relative to controls) were
observed in fish fed the highest dietary selenium treatment (12 mg Se/kg dw)
after the first 80 weeks of exposure just prior to transfer outdoors. No signs of
clinical disease were evident, and no relationship was found between feed
conversion ratios and the level of selenium added to the feed. Average whole
body selenium levels of female Henry's Lake cutthroat trout at spawning at 2.5 to
3 years of age were 5.87, 9.10, 11.37 and 5.61 mg Se/kg dw in the four highest
dietary treatments. Average egg selenium levels in the same four dietary
treatments were 6.61, 5.05, 5.18, and 16.04 mg Se/kg dw. Percent survival from
the eyed stage to hatching varied among treatment groups, with the control and
the highest Se dietary treatment having the second highest survival (85%) and the
fifth dietary treatment group the highest (93%). Percent deformed larvae ranged
from a low of 5.6% in controls to a high of 20.2% in the 6.4 mg Se/kg dw dietary
treatment group; larvae in the two highest dietary treatment groups only
exhibited 7 and 6.8 %, respectively.
Chronic Value:	The chronic value for embryo/larval deformity is aNOAEC of >11.37 mg Se/kg
dw whole-body parent tissue and >16.04 mg Se/kg dw egg.
C-52

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Rudolph, B-L, I. Andreller, CJ. Kennedy. 2008. Reproductive success, early life stage development,
and survival of Westslope cutthroat trout (Oncorhynchus clarkii lewisi) exposed to elevated selenium in
an area of active coal mining. Environ. Sci. Technol. 42: 3109-3114.
Test Organism:	Westslope cutthroat trout (Oncorhynchus clarkii lewisi)
Exposure Route: Field collected.
In June, 2005, eggs were collected from 12 females from Clode Pond (exposed
site) and 16 females from O'Rourke Lake (reference site). Milt was obtained
from 3-5 males at each site. Clode Pond is on the property of Fording River Coal
Operations in Southeast British Columbia with reported selenium concentrations
of 93 (ig/L. O'Rourke Lake is an isolated water body into which Westslope
cutthroat trout were stocked in 1985, 1989 and 1992 and has selenium levels
reported <1 |_ig/L.
Test duration:	Through the end of yolk sac absorption (at swim-up) by the alevins.
Study Design:	Individual batches of eggs were fertilized in the field with 2 ml composites of
milt. Water-hardened eggs were transported to the rearing laboratory. Eggs and
alevins were monitored daily for fertilization, hatching and mortality. After the
yolk sacs were absorbed, alevins were sacrificed and preserved in Davidson's
solution.
All viable fry (n = 4,922) after yolk absorption were observed for the frequency
and severity of skeletal (lordosis, kyphosis, and scoliosis), craniofacial (head,
eyes or jaw), and fin malformations as well as edema. The authors used a
graduated severity index (GSI) for deformities in which fry were scored 0
(normal) to 3 (severe) based on the level of defect.
Effects Data:	Eggs with the four highest Se concentrations (86.3 to 140 mg/kg dw) collected
from Clode Pond fish died before reaching the laboratory (Table 1). Excluding
the eggs that died from females CP1, CP3, CP4 and CP5, fertilization (total eggs
reaching the eyed stage/total eggs x 100) was not related to Se concentrations in
the eggs. The percent of alevins (post hatch to swim-up stage) that died was
related to the selenium concentration in the eggs (Table 1). Note: The data used
to estimate the ECi0 value excluded the variable from OL1 and OL2 (shaded
areas in Table 1). These are data from the reference lake in which only 57% of
the larvae survived (OL1) or where the % dead eggs plus % hatch did not add up
to % 100. Alevin survival was meaningfully higher in the other 15 clutches of
eggs from the reference site (85.1 to 99.8%). Because there were insufficient
partial effects, a TRAP model was not used to estimate the ECi0 value. The data
consist of a cluster background data and a cluster of 100% mortality (Figure 1).
With no way to fit a credible curve, the interpolation method is applied here with
the EC0 set to 20.6 mg/kg with background % survival of 95.75% (not including
the one low outlier) and the second extrapolation point being 46.8 mg/kg with
0.3% survival. The resultant slope is 5.6 (similar to slopes in other datasets where
it was estimated) and the ECio is 24.7 mg/kg. Note: TRAP was used in the
previous draft to derive a similar ECi0 of 24.1 mg/kg, however as stated above, it
C-53

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was determined that the data are not amenable to a TRAP model because of
insufficient partial effects.
An ECio based on Se in maternal muscle was estimated using the same approach
as was used for Se in eggs, that is, by interpolation between an EC0 and a high
ECP. An EC 10 of 16.6 mg Se/kg muscle dw was interpolated from an EC0
(HNOEC) of 13.4 mg/kg and the average background survival of 95.75 and the
ECioo set to 34.7 mg/kg muscle (Figure 2).
Deformity analysis was not performed on the alevins that died prior to the swim-
up stage. Therefore, due either to dead eggs or dead alevins, the occurrence and
severity of deformities were assessed on four clutches of eggs from Clode Pond
(CP2, CP6, CP11 and CP 12) with a range of 11.8 to 20.6 :g Se/g dw and 15 of
the 16 clutches (all eggs died in OL8) from O'Rourke Lake (Table 1). There was
no correlation between egg Se concentration and frequency of deformity or
edema. Statistical differences between sites were observed (p < 0.05) for skeletal
deformities and edema for both the frequency of the occurrence and the severity
score (Table 2). Note: the percent and severity score of skeletal deformities were
greater in the reference site than in the exposed site.
The effect level for this study was based on the alevin mortality data and not the
deformity measurements. Although edema occurred statistically more often at the
exposed site (87.7% at Clode Pond, 61.2% at O'Rourke Lake), it was not
correlated with selenium levels in the eggs. Also the greater occurrence of
skeletal malformations in the reference site confounded the use of statistical
differences between sites to determine effect levels for this study.
Effect Concentration: 24.7 mg Se/kg dw in eggs; 16.6 mg Se/kg dw in muscle.
C-54

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Table 1. Fertilization, egg mortality and alevin mortality for offspring from individual fish
collected in Clode Pond and O'Rourke Lake.

Muscle [Se]
Egg [Se]

Dead
Dead
%
Fish ID
mg/kg dw
mg/kg dw
Hatch %
eggs, %
alevins, %
Survivj
Clode Pond






(exposed site)






CP1
38.8
88.3
0
100
NA

CP2
11.8
16.1
98.2
1.8
0.9
99.1
CP3
40.4
86.3
0
100
NA

CP4
46.1
121
0
100
NA

CP5
50.4
140
0
100
NA

CP6
34.7
51
92.6
7.4
92.6
0.0
CP7
39
65.3
91.1
8.9
91.1
0.0
CP 8
7
11.8
63.9
36.1
0.8
98.7
CP9
35.4
46.8
63.4
36.6
63.2
0.3
CP10
35.5
75.4
82.4
17.6
82.4
0.0
cpii
11.3
16.9
77.9
22.1
1.3
98.3
CP12
13.4
20.6
97
3
5.1
94.7
avg
30.3
61.6
55.5
44
42
20.0
SD
15.1
42.4
42.5
42
44
0.0
O'Rourke Lake
(reference site)
Oil
s :x
i: <)
71.4
:x h
42 w
39.9
oi.:
7.7
13 w
27.7
53 I
<> <)
75.1
OL3
8.16
12.5
96.1
3.9
2.4
97.5
OL4
8.03
15
85.5
14.5
12.7
85.1
OL5
8.12
14.9
80.7
19.3
5.3
93.4
OL6
6.61
15.2
68
32
4
94.1
OL7
8.52
12.9
97.9
2.1
0.2
99.8
OL8
7.22
12.3
0
100
NA

OL9
7.25
16.7
87.2
12.8
4.5
94.8
OLIO
7.64
13.1
79.6
2.5
5.5
93.1
OL11
8.74
15.6
89.2
10.8
2.4
97.3
OL12
8.2
13.9
83.6
16.4
3
96.4
OL13
7.86
15.1
74.1
25.9
2.8
96.2
OL14
8.5
13.1
77.8
22.2
0.5
99.4
OL15
7.62
12.3
88.2
11.8
2.6
97.1
OL16
8.13
12.7
54.8
45.2
4.8
91.2
avg
7.9
13.9
72.6
25
7

SD
0.6
1.4
25.8
25
10

1 % Survival based on % hatch
C-55

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Table 2. Deformity results (frequency and severity) for offspring from O'Rourke Lake and Clode
Pond. Values are presented as mean  SE. * indicates a significant difference (p < 0.05) between means
from the two sites.
Frequency of deformity, %
O'Rourke Lake
Clode Pond
Skeletal*
37.4 3.6
16.5 2.2
Craniofacial
10.2 2.0
5.7  1.0
Finfold
10.6 3.1
7.5 3.84
Edema*
61.2 4.9
87.7 2.0
Severity of deformity, score


Skeletal*
0.47 0.07
0.18 0.02
Craniofacial
0.12 0.03
0.06 0.01
Finfold
0.15 0.05
0.09 0.05
Edema*
0.61 0.05
0.88 0.02
CD
>
w
100 -
80 -
60 -
40 -
20 -
10
20
50
1
100
mg Se/kg dw egg
Figure 1. Post-hatch survival of Westslope cutthroat trout alevin as a function of the logarithm of
the selenium concentration in eggs.
C-56

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100
80
<0
> 60
W
5- 40
20
i
10
i
20
50
mg Se/kg muscle dw
Figure 2. Post-hatch survival of Westslope cutthroat trout alevin as a function of the logarithm of
the selenium concentration in maternal muscle.
C-57

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Nautilus Environmental. 2011. Evaluation of the Effects of Selenium on Early Life Stage Development
of Westslope Cutthroat Trout from the Elk Valley, BC. Report to Elk Valley Selenium Task Force,
November 24, 2011.
Test Organism:	Westslope cutthroat trout (Oncorhynchus clarkii lewisi)
Exposure Route: Field collected. Adult fish were collected and spawned from lentic and lotic
environments in areas proximate to Teck Coal's Fording River Operations. Eggs
were also obtained from fish collected from Connor Lake, a lake located within
the Elk valley watershed not exposed to mine discharges and considered a
reference site and a methodological control.
Test Duration:	Fertilized eggs were reared in the laboratory until they reached swim-up fry
stage. A subset of fry surviving at swim-up were reared for an additional 28 days.
Study Design:	Gametes were stripped from the ripe adults in the field during June and July 2008
and transported immediately to the laboratory in coolers containing wet ice. Eggs
were fertilized in the laboratory. After stripping the eggs, female fish were
sacrificed and the whole body stored on ice for later Se analysis. For a given
female, approximately 240 fertilized eggs were divided into four replicates of 60
eggs. In cases when fewer eggs were available three replicates of 60 eggs were
used. If less than 180 eggs were available, either 3 or 4 replicates of 30 were
used. Females with less than 90 eggs were not used. The fertilized eggs were
maintained in the laboratory until the fry reached swim-up at which point
deformities were assessed. Survival was also assessed up to swim-up. In test
chambers in which there were at least 40 surviving fish at swim-up, one-half of
the surviving fish were maintained for an additional 28 days. Survival, length,
weight and deformities were assessed in the 28-day post swim-up test.
The number, type and severity of deformities were measured at swim-up and at
the end of the 28-day post swim-up test. Deformity assessments were conducted
on recently killed fresh fish to avoid artifacts caused by preservation. A
graduated severity index (GSI) was assigned to each of four types of
deformity/abnormality: skeletal, craniofacial, finfold and edema. Graduated
Severity Index (GSI) methods followed those described in Holm et al. (2003) and
Rudolph et al (2006; 2008).
Effects Data:	Survival of the larvae from hatch through swim-up spawned from the four fish
collected from the reference site, Connor Lake, ranged from 73 to 92% (egg Se
4.32 to 7.31 mg/kg dw) (Table 1). Larval survival at swim-up was also generally
high for fish collected in the Se exposed sites up to egg Se concentration 29.6
mg/kg dw (Table 1, Figure 1). Larvae exposed above this egg Se concentration
had poor to no survival. Larvae from one fish (P00811) below this threshold did
have poor survival (11.7%). The authors noted that the many of the eggs from
this fish displayed an unusual distribution of lipid vesicles which resulted in
greater than 50% mortality in the first 24 hours due to egg breakage. The
remaining eggs may have been compromised due to the organic material released
during the egg breakage.
C-58

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The rate of deformities in larvae at swim-up showed no relationship with Se in
egg through 29.6 mg/kg dw (Table 2).
The results of the 28-day post swim-up test showed no relationships between
larval survival or deformities and egg Se (Table 3). The authors also measured
the length and weight of larvae at the end of the 28 day test; neither of which
showed a relationship with egg Se concentration.
Se Tissue Concentrations. Two analytical laboratories (A and B) measured Se in
the eggs. The mean difference in egg Se concentrations between the two
laboratories was 34.2%. To better understand the difference between the two
laboratories, five egg samples (i.e., from five different fish) from this study were
sent to both laboratories in 2010. Both laboratories digested the eggs using the
methods they used in their own 2008 original analysis. The respective digestates
were split and then shared between laboratories. Both labs then measured
selenium in their own digestates and the digestate received from the other lab.
The results of this follow-up study showed that when each lab used their own
digestion procedures Laboratory A had on average 43% higher measurements in
the 2008 analysis and 23% higher in the follow-up 2010 analysis. When each lab
measured selenium using the same digestate the difference in the Se
measurements between labs was on average only 1 to 8%. The authors concluded
that although both laboratories employed acceptable and approved practices,
Laboratory A used a more efficient digestion process resulting in higher Se
measurements. To compensate for the reduced Se measurements in Laboratory B,
its values were increased by 34.2%. The measurements made by Laboratory A
are marked in Table 1; unmarked values are Laboratory B measurements
increased by 34.2%.
Effect Concentration: The most sensitive endpoint determined by TRAP was larval survival at swim-
up. Interpolation was used to estimate an effect concentration for larval survival
with the entire egg Se dataset that included egg Se measurements from
Laboratory A and adjusted measurements from Laboratory B (ECi0 = 31.1 mg/kg
egg dw; Figure 1) and using only the egg Se measurements from Laboratory A
(Figure 2). Because the Laboratory A dataset estimated slightly lower EC values,
the ECio of 27.7 mg/kg egg dw is the selected effect concentration for this study.
Note: In the previous draft, a TRAP model was used to estimate the ECi0.
However, because of insufficient partial effects, TRAP was determined not
appropriate so the ECi0 was estimated using an interpolation between the
HNOEC and the LOEC (see Figure 3 for the TRAP analysis used in the previous
draft).
C-59

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Table 1. Summary of westslope cutthroat trout larvae surviving to swim-up per parent female (fish
ID) including location of collection of parent female and concentration of selenium in the eggs.
	Proportion surviving	
Se egg,	Replicate Replicate Replicate Number Total
Fish ID
Location
mg/kg dw
Replicates
mean
min
max
survivors
number
Y093
Lentic
3.88*
4
0.8125
0.6667
0.9167
195
240
CL1
Reference
4.32
4
0.9167
0.8833
1
220
240
R082
Lotic
5.21
3
0.9056
0.8333
0.95
163
180
CL4
Reference
5.96*
4
0.7333
0.6
0.8
176
240
CL2
Reference
6.82
4
0.8333
0.7
0.9167
200
240
CL3
Reference
7.31
4
0.8542
0.8167
0.8833
205
240
P00815
Lotic
7.6
3
0.8222
0.7167
0.95
148
180
R026
Lotic
12.53
4
0.5792
0.5
0.65
139
240
P00823
Lotic
12.71
4
0.8875
0.85
0.95
213
240
R039
Lotic
12.9
4
0.6042
0.55
0.65
145
240
R086
Lotic
13.4*
4
0.9417
0.85
0.9833
226
240
R077
Lotic
14.29
3
0.6444
0.6167
0.6667
116
180
R042
Lotic
16.44
3
0.8
0.7
0.9
72
90
R055
Lotic
16.5
4
0.8792
0.7833
0.9667
211
240
R043
Lotic
16.85
4
0.8667
0.7667
0.9667
104
120
R074
Lotic
17.8*
4
0.9375
0.8833
0.9833
225
240
P00811
Lotic
19.25
1
0.1167
0.1167
0.1167
7
60
P00809
Lotic
19.72
4
0.7667
0.65
0.8833
184
240
P00803
Lotic
24.8*
4
0.9375
0.9333
0.95
225
240
R078
Lotic
29.61
4
0.8825
0.8333
0.9333
105
119
G099
Lotic
34.2*
4
0.2083
0.1667
0.2667
50
240
0087
Lentic
54.7*
4
0.07083
0.01667
0.2
17
240
0085
Lentic
56.8*
4
0
0
0
0
240
W052
Lentic
61.1*
4
0
0
0
0
240
R069
Lotic
65.61
4
0
0
0
0
240
R071
Lotic
72.9
4
0
0
0
0
240
W094
Lentic
73.1
4
0
0
0
0
240
UT101
Lentic
74.67
4
0
0
0
0
240
*Laboratory A dataset
C-60

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Table 2. Summary of westslope cutthroat trout larval deformities to swim-up per parent female
(fish ID) including location of collection of parent female and concentration of selenium in the eggs.
Fish ID
Location
Se egg,
mg/kg dw
Skeletal
combined
Craniofacial
combined
Finfold
combined
Edema
combined
Deformities
combined
Y093
Lentic
3.88*
4.5%
0.9%
4.4%
1.9%
7.7%
CL1
Lentic
4.32
7.6%
1.9%
1.0%
1.0%
9.5%
R082
Lotic
5.21
1.2%
1.3%
2.5%
0.0%
3.7%
CL4
Lentic
5.96*
4.3%
7.3%
1.7%
0.7%
12.6%
CL2
Lentic
6.82
11.1%
3.7%
0.8%
3.0%
15.9%
CL3
Lentic
7.31
5.0%
2.0%
1.0%
0.0%
7.0%
P00815
Lotic
7.6
0.0%
2.7%
0.0%
2.9%
5.6%
R026
Lotic
12.53
2.1%
2.1%
0.7%
1.4%
2.1%
P00823
Lotic
12.71
1.9%
2.9%
1.8%
5.6%
7.4%
R039
Lotic
12.9
2.1%
1.9%
2.9%
4.9%
9.9%
R086
Lotic
13.4*
2.7%
1.0%
0.0%
0.0%
2.7%
R077
Lotic
14.29
1.7%
10.4%
0.9%
12.2%
15.5%
R042
Lotic
16.44
1.2%
0.0%
0.0%
2.6%
2.6%
R055
Lotic
16.5
0.0%
2.8%
1.0%
2.9%
4.7%
R043
Lotic
16.85
0.9%
2.6%
1.8%
1.7%
4.4%
R074
Lotic
17.8*
2.7%
1.8%
0.9%
0.9%
3.6%
P00809
Lotic
19.72
3.9%
2.8%
3.3%
4.7%
9.0%
P00803
Lotic
24.8*
2.7%
0.9%
0.0%
0.9%
4.5%
G092
Lotic
26.1
0.0%
1.9%
1.9%
4.4%
4.4%
R078
Lotic
29.61
1.8%
0.0%
1.0%
2.9%
5.7%
G099
Lotic
34.2*
14.5%
53.9%
6.8%
28.2%
64.7%
*Laboratory A dataset
C-61

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Table 3. Summary of larval survival and rates deformities after the 28-day post swim-up test per
parent female (fish ID) including location of collection of parent female and concentration of
selenium in the eggs.	


Sample
Egg Se
Survival
Skeletal
Craniofacial
Finfold
Total
Fish ID
Location
size (n)
(mg/kg dw)
(%)
(%)
(%)
(%)
(%)
CL1
Reference
112
4.3
99.1
0
0
0
0
CL2
Reference
93
6.8
99
0
0
0
0
CL3
Reference
96
7.3
91.7
0
1
1
2
CL4
Reference
68
6
98.6
0
0
4.3
4.3
Y093
Lentic
93
3.9
95.6
0
0
2
2
R082
Lotic
71
5.2
87.4
0
2.9
0
2.9
P00815
Lotic
69
7.6
91.1
0
1.2
1.4
2
P00823
Lotic
105
12.7
96.3
0
0
0
0
R086
Lotic
112
13.4
97.2
0
0.9
0
0.9
R077
Lotic
36
14.3
92.4
2.8
2.8
2.8
4.2
R055
Lotic
101
16.5
95.9
0
4.6
0
4.6
R074
Lotic
106
17.8
93.1
0
0
0
0
P00809
Lotic
65
19.7
91.7
0
0
0
0
P00803
Lotic
108
24.8
95.7
0
0
1
1
100-
mg Se/kg egg
Figure 1. Labs A and B datasets. EC10 based on interpolation between the one partial effect (34.2
mg/kg, 20.8%) and an ECO set at the HNOEC and the average % survival for all the NOECs (29.6
mg/kg and 81.1%). The slope is 20.5 and the EC 10 is 31.1 mg/kg. Note: the gray point denotes egg
batch with quality problems noted by authors and was not used in the analysis.
C-62

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100 -
mg Se/kg egg
Figure 2. Lab A dataset.* EC10 based on interpolation between the one partial effect (34.2 mg/kg,
20.8%) and an ECO set at the HNOEC and the average % survival for all the NOECs (24.8 mg/kg
and 87.25%). The slope is 9.4 and the EC10 is 27.7 mg/kg.
*Although some scientists have attempted to explain certain occurrences of improved response with
increasing concentration in terms of nutrient selenium sufficiency-deficiency, the concentrations involved
in this study are too high to for selenium deficiency to be an explanation. The figure's apparent bi-phasic
measured response is thus best explained as being a chance outcome of noise.
C-63

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1.1
1.0
.9
.8
.7
.6
.5
.4
.3
.2
0
.4
.6
.8
1.0
1.2
1.4
1.6
1.8
2.0
Log(Se egg), mg/kg dw
Figure 3. (From previous draft) Tolerance distribution; Model option - Triangular distribution (3
parameter). Includes Laboratory "A" dataset only TRAP ECi0 estimate = 24.0 mg/kg.
C-64

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Golder Associates. 2009. Development of a Site-specific Selenium Toxicity Threshold for Dolly Varden
Char. Report to Northgate Minerals Corporation, PO Box 3519, Smithers, British Columbia. Report
Number 04-1421-101/2000.
Test Organism:	Dolly Varden (Salvelinus malma)
Exposure Route: Field collected.
Adult Dolly Varden char were collected from reference (North Kemess Creek),
high Se exposure (Upper Waste Rock Ponds and Creek) and moderate Se
exposure (lower Waste Rock Creek) sites during September 22 to 24, 2008. Eggs
were stripped from females and fertilized with milt from males collected from the
reference site. Fertilized eggs were taken to the laboratory for testing.
Test duration:	The test was terminated when 90% of the larvae reached swim-up, approximately
5 months after fertilization.
Study Design:	Approximately 30 fertilized eggs were added to each replicate rearing container.
The number of replicates per female parent ranged from one to four depending
on the number of eggs available. Embryos were maintained in 4 L containers
with 3.5 L dechlorinated tap water in a static-renewal system (3 renewals
times/week) at 5C. The condition of the embryos and alevins were observed
daily and any dead individuals were counted and removed. Test termination
occurred over a 3-day period during February 11 to 13, 2009. The hatched larvae
were sacrificed using an overdose of the anesthetic, clove oil. Individual length
and weight were measured on each fry, and deformity analysis was performed on
fresh unpreserved larval fish using 40X magnification.
A graduated severity index (GSI) was used for deformity assessment (skeletal,
craniofacial, and finfold as well as edema). The narrative criteria were the same
as used by Holm et al. (2005) and Rudolph et al. (2008).
Effects Data:	Alevin survival was not related to Se concentration in the eggs (Table 1). Almost
all of the mortality occurred during the egg stage. Only 4 alevins died during the
study, 1 from Fish #19 and 3 from Fish #2, both females collected at an exposed
site. The prevalence of deformities increased sharply after the selenium egg
concentration exceeded 50 mg/kg dw (Table 1, Figure 1). The proportion of
Dolly Varden larvae with any type of deformity (skeletal, craniofacial, and
finfold as well as edema) as a function of the log of the selenium concentration in
the eggs using TRAP (logistic equation) produced an ECi0 value of 56.22 mg/kg
dw eggs (Figure 1).
C-65

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Table 1. Selenium concentration in the eggs of Dolly Varden char and the survival of alevins to the
swim-up stage and the proportion of larvae without any type of deformity.
Survival of eggs to swim- Proportion of
[Se] 	up	 larvae
eggs	without any
Fish
#
Sample
ID
Location
mg/kg
dw
Initial
End
%
type of
deformity

WRC-






1
F105
Waste Rock Creek
56.6
120
71
59
0.89
2
WRC-F61
Waste Rock Creek
65.8
120
81
68
0.58

WRC-






5
F103
Waste Rock Creek
32.6
29
29
100
0.97
6
WRC-F83
Waste Rock Creek
51.9
120
115
96
0.97

WRC-






15
F104
Waste Rock Creek
56.3
60
48
80
0.90
19
WRC-F86
Waste Rock Creek
60.5
120
115
96
0.72
9
NK-F30
North Kemess Creek
11
30
1
3
a
12
NK-F29
North Kemess Creek
10.5
46
15
33
1.00
17
NK-F21
North Kemess Creek
5.4
90
86
96
0.91


Southern Collection





SCD1
Redd # 1
Ditch
10.3
30
18
60
1.00


Southern Collection





SCD2
Redd #2
Ditch
24.7
40
32
80
1.00
C-66

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Figure C-l. Proportion of Dolly Varden alevin without any type of deformity as a logistic
function of the logarithm of the selenium concentration in eggs (TRAP).
1.2 r
E 1.0
o
O
CD
CO
O
o
o
Q_
O
r =0.933
1.0	1.2	1.4	1.6
Iog10 Egg Se (mg/kg dw)
2.0

Guess
Final
SE
95% LCL
95% UCL
LogX50
1.844
1.829
0.007
1.812
1.845
Slope
4.152
6.963
1.252
4.003
9.924
YO
0.975
0.980
0.017
0.939
1.021

ECx
EC
95% LCL
95% UCL


50
67.42
64.92
70.01


20
60.12
57.96
62.35


10
56.22
53.00
59.64


5
52.85
48.64
57.43


1
46.11
40.13
52.99


DF
SS
MS
F
P
Total
9
1.74E-01
1.93E-02


Model
2
1.63E-01
8.13E-02
51.429
0.99993
Error
7
1.11E-02
1.58E-03


C-67

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AECOM. 2012. Reproductive success study with brown trout (Salmo trutta). Data quality assurance
report. Final. December 2012.
Formation Environmental. 2011. Brown Trout Laboratory Reproduction Studies Conducted in Support
of Development of a Site-Specific Selenium Criterion. Prepared for J.R. Simplot Company by Formation
Environmental. Revised October 2011.
Test Organism:	Brown trout (Salmo trutta)
Exposure Route: Field collected.
Adult female and male brown trout were collected at three field sites from two
streams downstream of the Smokey Canyon mine. In addition, brown trout eggs
were obtained from two hatcheries as method controls.
Test duration:	Embryo-larval monitoring to 15 days post swim-up.
Study Design:	Eggs were collected from 26 ripe female brown trout at three field sites
downstream of the Smokey Canyon mine. These included one site on the highly
impacted Sage Creek (LSV2C) as well as two sites along Crow Creek (CC-150
and CC-350) downstream of the conflux with Sage Creek. The downstream -
most station along Crow Creek (CC-150) was intended to be a field control. Eggs
were fertilized in the field with milt collected from males collected at the same
site as females. Fertilized eggs were water hardened at the site using stream
water, then placed in oxygenated plastic bags and stored on ice in the dark
(cooler) for transportation to laboratory. Selenium was measured in adult fish
(whole body) and in eggs of field collected females. In addition, eggs were
collected from 8 ripe females obtained from the Saratoga National Fish Hatchery
(SC) to serve as method controls. Similar to field-caught fish, SC hatchery
females were stripped of eggs and fertilized by milt from males obtained from
the same hatchery. As a result of lower than expected hatch rates and fungal
contamination in some SC hatchery samples, additional hatchery fish were
obtained (as already fertilized eyed embryos) from the Spring Creek Trout
Hatchery (SPC), which were divided into four treatments.
Approximately 600 fertilized eggs from each female (or 600 eyed embryos for
SPC treatments) were placed in egg cups for hatching and monitoring. After
swim up, remaining fry were thinned to a target of 100 fry/treatment and
monitored for an additional 15-day post swim up feeding trial. Test termination
ranged from 83 to 88 days after hatch for all but the Spring Creek Hatchery egg
treatments, which occurred 50 days after the arrival of fertilized, eyed embryos
from that hatchery.
Endpoints measured in the laboratory study were fecundity, hatch, growth,
survival/mortality, and feeding success (growth) post swim up. Larval brown
trout were also evaluated for deformities (craniofacial, vertebral, fin) and edema.
For this study, deformities were combined and assessed as having at least one
deformity, or being fully free of deformities (i.e., normal).
C-68

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Effects Data:	Se concentrations in eggs ranged from 6.2-12.8 mg Se/kg dw at CC150, 6.9-14.0
mg Se/kg dw at CC350, and 11.2-40.3 mg Se/kg dw at LSV2C. Se
concentrations in hatchery eggs ranged from 0.76-1.2 mg Se/kg dw at the SC
hatchery, and were 0.73 mg Se/kg dw at the SPC hatchery. The Se whole body
concentration in field collected fish ranged from 7.2-22.6 mg/kg dw at LSV2C,
4.7-8.4 mg/kg dw at CC150, and 5.5-9.2 mg/kg dw at CC350. Se whole body
concentrations in SC hatchery fish ranged from 2.5-4.3 mg/kg dw. Hatchery data
were combined with field data and included in all analyses.
Three endpoints were considered for purposes of calculating an ECi0. These were
percent survival, percent fully free from deformities, and percent surviving and
normal. Initially, data for these endpoints were combined and analyzed for both
portions of the test: hatch through swim up and the 15-day post swim feeding
trial. Data for these endpoints over both portions of the test are shown in Tables
1-3.
A U.S. Fish and Wildlife (2012) review of the Formation Environmental (2011)
report suggested that fish lost due to an overflow even resulting from a drain the
became clogged with food during the 15-day post swim up portion of the test
were more likely to have been dead or deformed, and proposed that all treatments
that lost fish to the overflow event should be excluded from the ECi0 calculation.
In the 2014 and 2015 draft Se documents, endpoints assessed for the hatch
through 15-day post swim up test were analyzed using two scenarios. In the
"worst-case" scenarios, the hypothesis from the USFWS review was examined,
by treating all fish lost to overflow as either dead or deformed, rather than
excluding those treatments altogether. In the "optimistic" scenario, the overflow
event was treated as a random technician error unrelated to selenium toxicity, and
any lost fish were removed from the calculation. In other words, fish lost to
overflow were assumed to be equally likely to have been dead or deformed
compared to fish that were not lost.
Because of the importance of these data for the numeric criterion calculation, and
because of several experimental factors that resulted in the calculation of several
reasonable ECi0s, such as the loss of fish due to an overflow event described
above, EPA conducted a careful and thorough reanalysis of the study data and
subjected the reanalysis to independent, external peer review (ERG 2012) to
confirm the validity and scientific robustness of the approach taken by EPA in
the reanalysis and use of the reanalyzed data. Those assessments were then
superseded by a reanalysis of a more complete enumeration of the deformity
counts provided by AECOM (2012). All analyses reported in the 2014 and 2015
draft Se documents and the current Se document used values from the updated
dataset provided by AECOM (2012).
Hatch Through 15-Day Post Swim Up Combined Data
In the 2014 and 2015 draft Se documents, data for three endpoints, survival,
deformities, and combined survival+deformities were considered for both
portions of the test. The first portion of the test was from hatch through swim up,
lasting 88 days (on average). The second portion was the 15-day post-swim up
C-69

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feeding trial. None of the fry from the five treatments with Se concentrations of
26.8 mg/kg and higher reached swim-up. However, surviving fry from those
treatments were included in the post-swim up feeding trial.
Combined Survival and Deformity Endpoint
Selenium concentrations and counts of total larvae, and counts of proportions of
fully normal larvae (alive and normal) are included in Table 1. Background
percentages of live and normal individuals were extremely variable and often low
(Figure 1). In the 2014 draft document, ECi0s for the optimistic (21.16 mg/kg)
and worst case (20.65 mg/kg) scenarios were calculated, and these were also
reported in the 2015 draft document. Although there is a clear demarcation
between treatments equal to or less than 20.5 (ig/L and treatments equal to or
greater than 26.8 (ig/L, suggesting an effect level between these concentrations, a
careful reanalysis of these data following the release of the 2015 draft Se
document determined that a meaningful ECi0 cannot be calculated because of the
high background variability.
C-70

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Table 1. Brown trout selenium concentrations and survival + deformity data (combined endpoint) from hatch to test end (15 days post
swim up).
Sample IDa
Whole
body
Se, mg/kg
d\v
Egg Se
mg/kg
dvv
#
Normal
# Normal
that were
dead at
assessment
#
Normal
and
alive
# Live fish
assessed for
deformities
# Fish died
during test
# Fish lost to
overflow
during post
swim up test
# Live fish
assessed + # died
during test.
Prop. Live fish
assessed + # died
during test.
SC-001
3.6
0.76
63

63
115
8

123
0.512
SC-002
4.1
0.94
72

72
113
4

117
0.615
SC-003
3.7
0.83
131

131
302
7
9
309
0.424
SC-004
4.3
0.92
46

46
140
28

168
0.274
SC-005
3
1.2
23

23
42
6

48
0.479
SC-006
3.1
1.2
457

457
535
8

543
0.842
SC-007
2.7
1
93

93
137
30

167
0.557
SC-008
2.5
0.96
283

283
359
6
10
365
0.775
SPC-001C

0.73
427

427
570
8

578
0.739
SPC-002C

0.73
371

371
545
20

565
0.657
SPC-005C

0.73
400

400
561
8

569
0.703
SPC-006C

0.73
427

427
556
17

573
0.745
CC-150-009
8.4
12.8
106

106
142
11

153
0.693
CC-150-011
5.6
8.4
87

87
266
2

268
0.325
CC-150-012
6.7
8.5
156

156
282
12

294
0.531
CC-150-013
5.9
8.4
137

137
310
46
26
356
0.385
CC-150-015
6
9.1
210

210
445
14

459
0.458
CC-150-016
7
7.5
13

13
23
3
43
26
0.500
CC-150-017
5.6
6.6
99

99
163
7
33
170
0.582
CC-150-018
4.7
6.9
195

195
486
16

502
0.388
CC-150-020
7.2
6.2
453

453
558
6

564
0.803
CC-350-006
9.2
14
120

120
386
26

412
0.291
CC-350-007
5.5
6.9
68

68
131
10
20
141
0.482
CC-350-008
8.5
9.5
269

269
338
21
28
359
0.749
LSV2C-002
8.9
12.8
483

483
544
4
16
548
0.881
LSV2C-003
13.8
40.3
2
2
0
0
395

395
0.000
LSV2C-004
17.9
36
16
16
0
0
289

289
0.000
LSV2C-005
13.6
26.8
8
8
0
0
267

267
0.000
LSV2C-008
9.6
17.7
147

147
194
4
45
198
0.742
C-71

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Whole


# Normal
#


# Fish lost to



body
Egg Se

that were
Normal
# Live fish

overflow
# Live fish
Prop. Live fish

Se, mg/kg
mg/kg
#
dead at
and
assessed for
# Fish died
during post
assessed + # died
assessed + # died
Sample ID11
dw
dw
Normal
assessment
alive
deformities
during test
swim up test
during test.
during test.
LSV2C-010
22.6
38.8
5
5
0
0
97

97
0.000
LSV2C-012
7.2
13.2
217

217
554
17

571
0.380
LSV2C-016
9.2
13.4
440

440
530
20

550
0.800
LSV2C-017
13.2
20.5
110

110
150
28
19
178
0.618
LSV2C-019
8.6
12.5
267

267
390
22
39
412
0.648
LSV2C-020
11.3
11.2
240

240
296
5
36
301
0.797
LSV2C-021
20
28.1
8
8
0
0
404

404
0.000
a SC - Saratoga National Fish Hatchery; SPC - Spring Creek Trout Hatchery; CC - Crow Creek; LSV - Sage Creek
b Test end was 15 days after swim up.
c Arrived as fertilized, eyed-eggs. No whole body Se measurement possible.
C-72

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SC Hatchery
^ SPC Hatchery
CC-150
CC-350
A LSV-2C
0.5	1.0
log(Dg Se/kg dw)
Figure 1. Proportion of alive and normal larvae plotted against Se
concentrations in eggs. Effects were highly variable across the entire
background concentration range (20.5 mg/kg and lower), such that a meaningful
ECio could not be calculated for this endpoint.
Deformity Endpoint
Selenium concentrations, counts of larvae assessed for deformities, and counts
and proportions of normal larvae are included in Table 2. As with the combined
endpoint, background (at or below 20.5 mg/kg) proportions of deformities were
highly variable (Figure 2). In the 2014 draft document, ECi0s were calculated for
both the optimistic and worst case scenarios, and the ECi0 of 15.91 mg/kg for the
worst case scenario was used as the ECi0 for Salmo. During the review phase
following the release of the 2014 draft Se document, several public commenters
noted that because of the high variability, more than one ECi0 could be calculated
by TRAP for both the optimistic and the worst case scenarios depending on the
initial model conditions, in particular the slope of the falling limb of the
concentration-response curve. For the optimistic scenario, ECi0s based on initial
conditions ranged from 16.36-21.95 mg/kg, and for the worst case scenario,
ECioS based on initial conditions ranged from 15.91-21.58 mg/kg. In order to
evaluate the most appropriate ECio for the deformity endpoints, models were
evaluated based on residual sum of squares, and the ECi0 for the model with the
lowest residual sum of squares was selected as the most appropriate. For the
worst case scenario deformity endpoint, the model with the lowest residual sum
C-73

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of squares was the ECi0=21.58 mg/kg model, and for the optimistic deformity
endpoint, the model with the lowest residual sum of squares was the ECi0=21.94
mg/kg model.
These variable ECi0s were the result of large variability in background
concentration, with several treatments at low Se concentrations experiencing
greater than 60% deformities (Figure 2). Although there is clear evidence of an
effect between the 20.5 and 26.8 mg/kg concentrations, because of this high
background variability, a careful re-analysis of these data following the release of
the 2015 draft Se document determined that a meaningful ECio could be
calculated for the deformity endpoint.
Some of the background variability in deformities appears to be the result of
differences among field sites. For example, deformity rates among field samples
appear to be greater for fish hatched from eggs collected in the two Crow Creek
sites (CC-150, CC-350) compared to Sage Creek (LSV-2C) (Figure 2). If the
result of higher background deformities among Crow Creek sites is not a random
artifact, it suggests a confounding factor, unrelated to selenium exposure.
Whether the higher deformity rates represent random variation, population
differences, other environmental quality differences (unrelated to Se), or
methodological issues is unclear.
Table 2. Brown trout selenium concentrations and deformity data from hatch to test end (15 days
post swim up).

Whole


# Assessed for
# Lost to


body
Egg Se

deformities.
overflow
Prop. Assessed
Sample
Se, mg/kg
mg/kg
#
"Optimistic
during post
for deformities
IDa
dw
dw
Normal
Case"
swim up test
plus # lost.
SC-001
3.6
0.76
63
115

0.548
SC-002
4.1
0.94
72
113

0.637
SC-003
3.7
0.83
131
302
9
0.434
SC-004
4.3
0.92
46
140

0.329
SC-005
3
1.2
23
42

0.548
SC-006
3.1
1.2
457
535

0.854
SC-007
2.7
1
93
137

0.679
SC-008
2.5
0.96
283
359
10
0.788
spc-ooic

0.73
427
570

0.749
SPC-002C

0.73
371
545

0.681
SPC-005C

0.73
400
561

0.713
SPC-006C

0.73
427
556

0.768
CC-150-





0.746
009
8.4
12.8
106
142

CC-150-





0.327
011
5.6
8.4
87
266

CC-150-





0.553
012
6.7
8.5
156
282

CC-150-





0.442
013
5.9
8.4
137
310
26
CC-150-
6
9.1
210
445

0.472
C-74

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Sample
IDa
015
CC-150-
016
CC-150-
017
CC-150-
018
CC-150-
020
CC-350-
006
CC-350-
007
CC-350-
008
LSV2C-
002
LSV2C-
003
LSV2C-
004
LSV2C-
005
LSV2C-
008
LSV2C-
010
LSV2C-
012
LSV2C-
016
LSV2C-
017
LSV2C-
019
LSV2C-
020
LSV2C-
021
Whole
body
Se, mg/kg
dw
7
5.6
4.7
7.2
9.2
5.5
8.5
8.9
13.8
17.9
13.6
9.6
22.6
7.2
9.2
13.2
8.6
11.3
20
Egg Se
mg/kg
dw
7.5
6.6
6.9
6.2
14
6.9
9.5
12.8
40.3
36
26.8
17.7
38.8
13.2
13.4
20.5
12.5
11.2
28.1
#
Normal
13
99
195
453
120
68
269
483
2
16
8
147
5
217
440
110
267
240
# Assessed for
deformities.
"Optimistic
Case"
23
163
486
558
386
131
338
544
100
142
149
194
80
554
530
150
390
296
172
# Lost to
overflow
during post
swim up test
43
33
20
28
16
45
19
39
36
Prop. Assessed
for deformities
plus # lost.
0.565
0.607
0.401
0.812
0.311
0.519
0.796
0.888
0.020
0.113
0.054
0.758
0.063
0.392
0.830
0.733
0.685
0.811
0.047
a SC - Saratoga National Fish Hatchery; SPC - Spring Creek Trout Hatchery; CC - Crow Creek; LSV -
Sage Creek
b Test end was 15 days after swim up.
c Arrived as fertilized, eyed-eggs. No whole body Se measurement possible.
C-75

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1.0
0.8 -
TO
E
i_
o
z 0.6 -
"5
u_
c
o
e 0.4 -
o
o.
o
i_
CL
0.2 -
0.0 -
0.0	0.5	1.0	1.5
log(Dg Se/kg dw)
Figure 2. Proportion of normal (free from deformities) larvae plotted
against Se concentrations in eggs, hatch through 15-days post swim up.
Effects were highly variable across the entire background concentration range
(20.5 mg/kg and lower), such that a meaningful ECi0 could not be calculated for
this endpoint.
Survival Endpoint
Selenium concentrations and estimated counts and proportions of larvae
surviving from hatch through 15 days post swim up are included in Table 3.
Estimated counts and proportions were reported for survival through the 15-day
post swim up test because larvae were thinned to a target of 100
individuals/treatment prior to the onset of the post swim up test, and final full test
survival is calculated as the product of survival from hatch to swim up and
survival during the 15-day post swim up test. In the 2014 draft document, ECi0s
were calculated for the worst case (16.78 mg/kg) and optimistic (20.40 mg/kg)
survival scenarios, and these were also reported in the 2015 draft document. For
both scenarios, the assumption was made that fry that failed to swim up would
not have survived, and so the survival for the post swim up portion of the test in
the 5 treatments with the highest selenium concentrations (26.8 mg/kg and
above) was set to zero. The ECi0 of 16.78 mg/kg for the optimistic is nearly
identical to the ECi0 for the worst case survival scenario of 16.76 mg/kg
presented in the response to the FWS review of the Formation Environmental
study (Taulbee et al. 2012), peer reviewed by ERG (2012).
In contrast to the deformity and combined deformity+survival endpoints,
background survival (concentrations up to and including 20.5 mg/kg) was much
less variable. Despite the lower variability among background effect levels, a
careful re-examination of these data following the release of the 2015 draft Se
T
	SC Hatchery
^	SPC Hatchery
Q	CC-150
^	CC-350
A	LSV-2C
-I-
A
A
A
A
C-76

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document determined that a meaningful ECio cannot be calculated by TRAP so
long as the assumption is made that fry failing to reach swim up are assumed to
be dead. This is because TRAP requires at least 2 partial effects to calculate an
ECio, and this dataset has no partial effects, but rather, a background range with
high and relatively stable survival through 20.5 mg/kg, and then no survival at
concentrations of 26.8 mg/kg and above (Figure 3). In order to calculate an ECio
for survival, the assumption regarding fry that failed to swim up was removed. In
addition, in order to remove the uncertainty introduced by the clogged drain
leading to the overflow and loss of fish from some of the treatments in the post
swim up test, the ECio for larval survival was calculated for the much longer
hatch through swim up portion of the test, as described below.
Table 3. Brown trout selenium concentrations and survival data from hatch to test end (15 days
post swim up).
Sample IDa
Whole body
Se, mg/kg dw
Egg Se
mg/kg dw
# Eggs
Hatched
Prop.
Survival.
Hatch to swim
up
Prop survival.
Post swim
up."
Prop survival.
Hatch to endb.
SC-001
3.6
0.76
144
0.951
0.990
0.942
SC-002
4.1
0.94
138
0.978
0.990
0.968
SC-003
3.7
0.83
340
0.982
0.989
0.971
SC-004
4.3
0.92
189
0.868
0.971
0.842
SC-005
3
1.2
70
0.914
1.000
0.914
SC-006
3.1
1.2
564
0.988
0.990
0.978
SC-007
2.7
1
188
0.856
0.970
0.830
SC-008
2.5
0.96
396
0.985
1.000
0.985
SPC-001C

0.73
598
0.987
1.000
0.987
SPC-002C

0.73
20
1.000
1.000
1.000
SPC-003C

0.73
585
0.966
1.000
0.966
SPC-004C

0.73
21
1.000
1.000
1.000
SPC-005C

0.73
589
0.986
1.000
0.986
SPC-006C

0.73
593
0.971
1.000
0.971
CC-150-009
8.4
12.8
173
0.942
0.990
0.933
CC-150-011
5.6
8.4
288
0.993
1.000
0.993
CC-150-012
6.7
8.5
314
0.965
0.990
0.955
CC-150-013
5.9
8.4
402
0.891
0.973
0.866
CC-150-015
6
9.1
479
0.971
1.000
0.971
CC-150-016
7
7.5
89
0.966
1.000
0.966
CC-150-017
5.6
6.6
223
0.969
1.000
0.969
CC-150-018
4.7
6.9
522
0.969
1.000
0.969
CC-150-020
7.2
6.2
584
0.990
1.000
0.990
CC-350-006
9.2
14
432
0.944
0.980
0.926
CC-350-007
5.5
6.9
181
0.950
0.988
0.938
CC-350-008
8.5
9.5
407
0.951
0.986
0.938
LSV2C-002
8.9
12.8
584
0.993
1.000
0.993
LSV2C-003d
13.8
40.3
404
0.079
0.281
0.022
LSV2C-004d
17.9
36
309
0.414
0.477
0.197
LSV2C-005d
13.6
26.8
287
0.387
0.622
0.240
LSV2C-008
9.6
17.7
263
0.989
0.982
0.971
LSV2C-010d
22.6
38.8
108
0.231
0.440
0.102
LSV2C-012
7.2
13.2
591
0.971
1.000
0.971
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Prop.
Survival.
Prop survival.

Whole body
Egg Se
# Eggs
Hatch to swim
Post swim
Prop survival.
Sample IDa
Se, mg/kg dw
mg/kg dw
Hatched
up
up."
Hatch to endb.
LSV2C-016
9.2
13.4
570
0.965
1.000
0.965
LSV2C-017
13.2
20.5
217
0.885
0.963
0.852
LSV2C-019
8.6
12.5
471
0.953
1.000
0.953
LSV2C-020
11.3
11.2
357
0.986
1.000
0.986
LSV2C-021d
20
28.1
424
0.288
0.730
0.210
a SC - Saratoga National Fish Hatchery; SPC - Spring Creek Trout Hatchery; CC - Crow Creek; LSV -
Sage Creek
b Test end was 15 days after swim up.
c Arrived as fertilized, eyed-eggs. No whole body Se measurement possible.d Survived but failed to reach
swim up. Assumed dead in all hatch to 15-day post swim up analysis.
1.2
1.0 -
O)
.E 0.8
>
'
3
(/)
C
o
'-e
o
Q_
O
0.6 -
0.4 -
0.2 -
0.0
w;
A

SC Hatchery
~
SPC Hatchery

CC-150

CC-350
A
LSV-2C
i
0.0
I
0.5
1.0
1.5
log(Dg Se/kg dw)
Figure 3. Proportion of larval survival plotted against log transformed Se
concentrations in eggs, hatch through 15-day post swim up. Larvae from the
five highest Se concentration treatments failed to reach swim up and were
assumed to not have survived in the wild.
Assessment of Overflow Loss During 15-dav Post Swim Up Feeding Trial
In the 2015 draft Se document, an assessment was made to determine whether the
loss of fish from the overflow event during the 15-day post swim up portion of
the test was related to survival or to Se treatment concentration measured during
the first portion of the test. In this assessment, data were examined from the
perspective of whether the overflow loss of brown trout during the second stage
of the test could reflect dead, dying, or weak organisms. This was done to
examine the hypothesis proposed in the U.S. FWS review that fish lost to
overflow were either dead or dying.
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First, the relationship between larval survival in the first and second stages of the
test (hatch to swim up, 15 days post swim up) were compared for all treatments
where larvae successfully reached the swim up stage (Figure 4). Overall, survival
in the second stage tracks survival in the first stage (r2=0.6), but survival in the
second stage was noticeably higher in than in the first stage. This result is
consistent with the following statement made by the principle scientist of the
brown trout study in the public comments to the 2014 selenium draft document
submitted for external peer review: "escaped fry were observed swimming in the
water bath where the treatment containers were being held. These fry
congregated near the treatment cells. Dead or dying fish were not observed."
m
bo
1.005
1
0.995
0.99
0.985
0.98
"1 0.975
0.97
0.965
0.96
$
s
I
3
E
p
&
f
0.84 0.86
0.88 0.9 0.92 0.94 0.96 0.98
Brown trout survival during 1st stage
1.02
Figure 4. Relationship between survival during the first and second portions
of the test. All treatments where larvae successfully reached swim up (Se
concentrations of 20.5 mg/kg and lower).
Second, the relationship between larval mortality in the first stage and overflow
loss in the second stages of the test (hatch to swim up, 15 days post swim up)
were compared separately for all treatments (field and hatchery) and for all field
collected treatments (Figure 5). As with figure 4, these correlations were made
for treatments where larvae successfully reached the swim up stage. In these
instances, there is no apparent relationship between health, as reflected by
mortality in the first stage, and overflow loss in the second stage, whether
considering all individuals or wild-only: r2 for both graphs is 0.0. The lack of a
relationship in these correlations suggests that overflow loss has a likelihood of
being a random noise variable.
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s.
%
0.6
0.5
oj 0.'
W>
0.3
0.2.
0.1
0
0.02 0.04 0.06 0.08 0.1 0.12
Brown trout mortality during 1st stage
0.14
0.16
0.6
8.
e
0.5
eM 0.'
(30
0.3
0.2.
0.1
0  #
0	0.02 0.04 0.06 0.08 0.1 0.12
Brown trout mortality during 1st stage
0.14
Figure 5. Relationship between mortality during the first stage of the test
and overflow loss during the second stage of the test. Upper figure - all
hatchery and field treatments. Lower figure - field treatments only. Larvae from
treatment levels 26.8 mg/kg and higher, which failed to swim up, were excluded.
Finally, the relationship between overflow loss and selenium concentrations in
eggs was examined (Figure 6). As with previous correlations, only larvae from
treatments where individuals reached swim up were considered.
Figure 6 shows a clear difference between hatchery (far left) and field treatments,
but across the concentration range for the offspring of field collected fish there is
no apparent relationship between overflow loss and Se concentration. Within the
field treatments, the r2 of the correlation between Se concentration and overflow
loss is 0.01. Although there are no known genetic differences between hatchery
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and wild fish, if leaving the aquarium required swimming over the rim, one
might speculate that previous generations of hatchery fish might have developed
a tolerance to remaining in conditions that might seem crowded to wild
organisms. (That is, however, purely speculative.) Otherwise, the difference
between hatchery and wild fish would seem only to reflect a random artifact,
since the Se concentrations at which the wild fish displayed high overflow losses
are low.
>
0.6
0,5
S 0.4
00
c
c
% 0.3
0.2
0.1
0
10	15
Egg Concentration, mg Se/kg dw
20
25
Figure 6. Relationship between egg Se concentration and overflow loss
during the second stage of the test. Larvae from treatment levels 26.8 mg/kg
and higher, which failed to swim up, were excluded.
In summary, the positive correlation between survival during the hatch to swim
up portion of the test and survival during the 15-day post swim up portion of the
test, combined with the lack of a correlation between mortality during the hatch
to swim up portion of the test and overflow loss during the second stage of the
test, suggests that the overflow loss likely represents a random technician error
not related to the health of the individuals lost. The relationship between
selenium egg concentrations and overflow loss was lower for the larvae hatched
from hatchery fish compared to the larvae hatched from field collected fish;
however, among field treatments ranging from 6.0-20.5 mg/kg there was no
correlation, further supporting the hypothesis that the overflow event was a
random occurrence unrelated to the health of larval fish.
The results of the above assessment of the overflow event strongly suggest that
the overflow event was a random technician error unrelated to selenium toxicity,
and that the "optimistic" scenario is also likely more realistic.
Survival Endpoint - ECm for the first portion of the test
Because larval survival was measured at the end of the first portion of the test
(hatch to swim up), an alternative approach to measuring survival would be to
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calculate the brown trout ECio for survival for only the first portion of the test.
Selenium concentrations and counts of total larvae and larvae that survived the
first portion of the test are included in Table 4. The hatch to swim up portion of
the test was much longer than the second portion (88 days on average compared
to 15 days), and more importantly, it avoids the experimental confound
introduced by the loss of fish during the overflow event. With this approach, the
second portion of the test would be rejected as inconclusive due to the laboratory
accident.
Unlike survival, deformities could not be analyzed for the first portion of the test
because of a bias introduced during the thinning process prior to the initiation of
the 15-day post swim up portion of the test. During the thinning process, visibly
deformed larvae were selectively removed, so