TECHNICAL MEMORANDUM 3: 

NORTH SAN FRANCISCO BAY SELENIUM 

TOXICOLOGICAL ASSESSMENT

DRAFT 

April 2008 

Prepared for 

Regional Water Quality Control Board 

San Francisco Bay Region 

1515 Clay Street 

Oakland, CA 94612 

Prepared by 

Tetra Tech, Inc. 

3746 Mt. Diablo Blvd., Suite 300 

Lafayette, CA 94549 

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TABLE OF CONTENTS

Abbreviations .............................................................................................................................. vii

1 Introduction .......................................................................................................................1-1

2 Invertebrate toxicity review................................................................................................2-1

2.1 Discussion................................................................................................................2-4

2.2 Conclusions..............................................................................................................2-4

3 Fish toxicity review ............................................................................................................3-1

3.1 Methods....................................................................................................................3-2

3.2 Results .....................................................................................................................3-4

3.3 Discussion..............................................................................................................3-13

3.4 Conclusions............................................................................................................3-18

4 Avian toxicity review..........................................................................................................4-1

4.1 Methods....................................................................................................................4-3

4.2 Results .....................................................................................................................4-5

4.3 Improved performance studies.................................................................................4-9

4.4 Discussion................................................................................................................4-9

4.5 Conclusions............................................................................................................4-18

5 Marine mammal toxicity review .........................................................................................5-1

5.1 Methods....................................................................................................................5-1

5.2 Results .....................................................................................................................5-1

5.3 Discussion................................................................................................................5-2

5.4 Conclusions..............................................................................................................5-5

6 Issues Affecting The Choice of Screening Values ............................................................6-1

7 Issues Affecting Implementation .......................................................................................7-1

8 References........................................................................................................................8-1

Appendix A Wet Weight to Dry Weight Conversions ................................................................ A-1

Appendix B Threshold Values for Selenium in Great Salt Lake................................................ B-1

Appendix C Allometric Adjustment of Dietary Concentrations .................................................. C-1

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LIST OF FIGURES

Figure 1-1 Ranked freshwater genus mean acute values for copper ..................................1-4

Figure 1-2 Health vs. dietary selenium concentration..........................................................1-5

Figure 1-3 Weight gain in fingerling channel catfish as a function of dietary selenium.......1-5

Figure 3-1 Biphasic regression analysis of the data from Hamilton et al. (1990).................3-5

Figure 3-2 Biphasic regression based on simulated data from Hamilton et al. (1990) ........3-6

Figure 3-3 Effect thresholds for centrarchids from the selected studies grouped by life 

history stage. ....................................................................................................3-11

Figure 3-4 Effect thresholds for rainbow trout from the selected studies grouped by life 

history stage. ....................................................................................................3-11

Figure 3-5 Effect thresholds for centrarchids after relaxing the selection criteria. .............3-12

Figure 3-6 Effect thresholds for rainbow trout after relaxing the selection criteria.............3-13

Figure 4-1 Logistic regression of mallard dietary selenium concentration vs. proportion 

of eggs hatching. ..............................................................................................4-11

Figure 4-2 Hockey stick regression of mallard dietary selenium concentration vs. 

proportion of eggs hatching..............................................................................4-12

Figure 4-3 Logistic regression of mallard egg selenium concentration vs. proportion 

of eggs hatching. ..............................................................................................4-12

Figure 4-4 Hockey stick regression of mallard egg selenium concentration vs. 

proportion of eggs hatching..............................................................................4-13

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LIST OF TABLES

Table 2-1 Summary of Results from Hyne et al. (2002) .....................................................2-3

Table 2-2 Effect thresholds for invertebrates exposed to dietary or sediment-bound 

selenium .............................................................................................................2-4

Table 3-1 Selenium toxicity studies evaluated in this review..............................................3-3

Table 3-2 The EC10 and EC20 from 10 simulated experiments and 95% confidence 

limits (CLs) on salmon survival at the EC10 and EC20......................................3-7

Table 3-3 Summary of selenium toxicity studies evaluated in this review..........................3-9

Table 3-4 Species mean chronic values calculated from the toxicity studies evaluated 

above................................................................................................................3-10

Table 3-5 Whole body fish tissue screening levels (mg/kg-dw) published in the 

scientific literature.............................................................................................3-14

Table 4-1 Selenium toxicity studies evaluated in this review..............................................4-5

Table 4-2 Summary of dietary selenium toxicity studies evaluated in this review. .............4-7

Table 4-3 Summary of additional selenium toxicity studies from the NRC (1980)..............4-8

Table 4-4 Summary of additional selenium toxicity studies from the NRC (1994)..............4-8

Table 4-5 Summary of additional selenium toxicity studies from the NRC (2005)..............4-8

Table 4-6 Studies evaluating improved performance in poultry from selenium 

supplementation. ................................................................................................4-9

Table 4-7 Summary of NOAELs and LOAELs (mg/kg-dw) reported in this section 

grouped by whether the effects were measured in chicks/ducklings, adult 

birds, or in terms of reproductive success........................................................4-10

Table 4-8 Selenium screening values for teratogenesis in bird eggs . .............................4-14

Table 4-9 Allometrically adjusted dietary selenium screening values...............................4-15

Table 4-10 Selenium nutritional requirements for commercially reared birds.....................4-15

Table 5-1 Dietary toxicity studies on pigs from NRC (1980, 2005).....................................5-2

Table 5-2 Dietary toxicity studies on dogs from NRC (1980)..............................................5-2

Table 5-3 Selenium nutritional requirements for commercially reared pigs by body 

weight .................................................................................................................5-3

Table C-1 Body weights and ingestion rates...................................................................... C-3

Table C-2 Allometric adjust factors and allometrically adjusted dietary thresholds............ C-3

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ABBREVIATIONS

AET apparent effects threshold 

ANOVA Analysis of variance 

CDH California Department of Health 

DFG Department of Fish and Game 

dw dry weight 

EC01 concentration at which an effect is observed in 1% of the population (i.e., 

effect concentration 1%) 

EC10 concentration at which an effect is observed in 10% of the population (i.e., 

effect concentration 10%) 

LC50 concentration lethal to 50% of test subjects 

LOAEL lowest observed adverse effects level 

MBTA Migratory Bird Treaty Act 

Na2SeO3 sodium selenite 

NRC National Research Council 

NIWQP National Irrigation Water Quality Control Program 

NOAEL no observed adverse effect level 

Se selenium 

SeMet selenomethionine 

SFEI San Francisco Estuary Institute 

SMCV Species mean chronic values are calculated by taking the geometric mean of 

the effect thresholds (e.g., NOAEL and LOAEL) for a particular species. 

TMDL total maximum daily load 

UCL95 95% upper confidence limit on the mean 

USDOI U.S. Department of the Interior 

USEPA U.S. Environmental Protection Agency 

USFWS U.S. Fish and Wildlife Service 
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USGS U.S. Geological Survey 

ww wet weight 
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1 INTRODUCTION

This technical memorandum provides a review of the available information on the toxicity 

of selenium to invertebrates, fish, birds, and marine mammals. This information has been 

developed to support the Regional Water Quality Control Board, San Francisco Bay Region 

(Water Board) in determining numeric targets for the selenium TMDL for the North San 

Francisco Bay (North Bay). The review of the selenium toxicity data was conducted in 

several steps. First, the available toxicity data from the published scientific literature and 

unpublished reports were compiled. These documents were identified from on-line searches 

of scientific references and the literature-cited sections of these reports were reviewed to 

identify unpublished reports. We estimate that greater than 95% of the relevant Englishlanguage literature has been reviewed. Next, the reported toxicity values were screened to 

identify those data that were most relevant to setting a numeric target for North Bay fish and 

wildlife. Screening criteria included: 

! Chronic exposure to single chemical (selenium) 

! Controlled experimental conditions 

! Dietary exposure 

! Tissue concentrations reported for exposed animals 

The reported toxicity values (tissue concentrations) for individual species can exhibit a wide 

range due to differences in the corresponding toxicity endpoint of the test. Therefore, the test 

endpoint merits special attention in interpreting the reported toxicity values. The tables in 

this report, summarizing the selenium toxicity studies evaluated, report both the toxicity 

value and the test endpoint. 

The focus of this report is on the compilation of relevant toxicity values. Some of the key 

issues to consider in interpreting the results of the laboratory tests and data analyses are also 

identified. 

1.1 BACKGROUND INFORMATION

The current EPA and California state freshwater water chronic quality criterion for the 

protection of aquatic life (5 μg/L) has been criticized as inadequately representative of the 

toxic risks of selenium to aquatic life. Recent studies on selenium toxicity in aquatic food 

chains have generally reached the conclusion that a water-based criterion is not suitable due 

to “...temporal [and spatial] changes in concentrations, speciation, and rates of transfer 

between water, sediment and organisms...” (Hamilton 2004). Since the primary route of 

exposure to selenium is via the diet, and selenium is highly bioaccumulative, these 

differences can mean that a concentration of selenium in water that results in adverse effects 

in one location may not result in adverse effects to the same species in another location. 

Thus, the current recommendation (USEPA 2004, Chapman 2007, Hamilton 2002, 2004) for 

the appropriate media for regulation of selenium in the aquatic environment is not water, but 

rather tissue. 
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In 1987 in the North San Francisco Bay (hereinafter referred to as the “North Bay”), “...the 

California Department of Health (CDH) issued a health advisory for the consumption of 

three [sic] species of diving ducks in Suisun Bay, near San Francisco. These diving ducks 

(i.e., surf scoter, lesser and greater scaup) feed on bottom-dwelling animals (clams, 

mussels), which concentrate selenium from sediment and phytoplankton. In a study by the 

California Department of Fish and Game, the mean selenium levels in the muscle of duck 

were as follows: scoters, 3.6 ppm Se, wet weight; scaups, 2.2 ppm Se, wet weight. These 

levels were up to 5 times higher than mean Se levels in similar species in control areas (i.e., 

Humboldt Bay near the border of Oregon and California). The CDH recommended limiting 

the consumption of scoters to 4 oz/2 wk and of scaups to 4 oz/wk. Because of the known 

effects of selenium on the reproduction of birds, the CDH recommended that women of 

child-bearing age and children, 15 years old not consume these ducks” (Barceloux 1999). 

In addition to the CDH health advisories for the consumption of diving ducks, the North Bay 

including Suisun Bay, Carquinez Strait, San Pablo Bay and Central Bay were listed in 1998, 

2002, and 2006 under section 303(d) of the Clean Water Act due to impairments to both 

wildlife and human-health beneficial uses (Abu-Saba and Ogle 2002). It is on the basis of 

these listings that the Regional Water Quality Control Board, San Francisco Region (Water 

Board) is developing the Total Maximum Daily Load for the North Bay. 

1.2 RELEVANT TOXICITY DATA

Several approaches have been used to develop screening values in the scientific and 

regulatory literature. These approaches include: 

! LOAELs 

! Effect thresholds 

! Species mean chronic values 

! EC01 or EC10 

! Species sensitivity distributions 

In the literature on selenium toxicity to fish, the approach that has generally been used is to 

set the screening value equal to the lowest LOAEL (i.e., the lowest observed adverse effect 

level) that has been reported (e.g., Hamilton 2003, 2004). When there is a large body of 

literature, with many reported LOAELs, this approach is likely to come close to the 

concentration at which effects first appear. However, when there are only a few studies, it is 

likely that effects begin at a level below the lowest LOAEL reported. For that reason, it may 

be advisable to use an SMCV, EC10, or effect threshold (see below) instead. 

Effect thresholds are calculated as the geometric mean of the NOAEL (i.e., the highest no

observed adverse effect level) and LOAEL reported for the same effect in an individual 

study. Since toxicity tests do not generally test many different concentrations, and effects 

may occur at concentrations below the LOAEL, calculating the geometric mean of the 

NOAEL and the LOAEL is one way to add a margin of safety to the LOAEL. That is, 
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effects are likely to occur somewhere between the NOAEL and the LOAEL. The USEPA 

(1985, 2003) has made the decision that the geometric mean of the NOAEL and LOAEL is 

representative of where those effects are likely to occur. 

Species mean chronic values (SMCVs) are used by USEPA in the calculation of water 

quality criteria (USEPA 1985). An SMCV is calculated by taking the geometric mean of the 

effect thresholds for a particular species. When there is only one study available, the results 

of that study are regarded as the SMCV. 

Another approach that can be used to find the concentration at which effects are likely to 

occur between the NOAEL and LOAEL is to pool the data from similar studies and perform 

a regression (e.g., biphasic, hockey-stick, or logistic) on the response data. The data is 

generally only from a single species. This approach has been used in State of Utah 

(Ohlendorf 2007) by pooling the data from 6 studies on mallards and performing regressions 

on the concentration of selenium in feed and eggs vs. reproductive success. The State of 

Utah derived an EC10 (i.e., the effect concentration 10% or the concentration at which an 

effect is observed in 10% of the population) for selenium in mallard diets and eggs 

protective of reproductive success. This same approach can also be used to calculate an 

EC01 (i.e., effect concentration 1%), which will be lower, but protective of a greater 

proportion of individuals. 

Species sensitivity distributions are another approach that have been used to derive 

screening values. In this approach, the probability distribution of some measure of toxicity 

(e.g., NOAEL, LOAEL, effect threshold, SMCV, LC50, etc.) for multiple animal species is 

calculated. From the probability distribution, a screening value is then calculated that is 

protective of a certain proportion of the species (e.g., 95%). USEPA calculates percentiles of 

species sensitivity distributions for use in the derivation of water quality criteria (Posthuma 

et al. 2002). This approach is similar to the EC10 approach described above, except that it 

always uses data from multiple species, with each data point representative of a single 

species, and uses probability distributions instead of regressions to derive screening values. 

For the recent acute criterion for copper, USEPA compared their calculated value to a 

species sensitivity distribution (Figure 1-1) and found that it was protective of 95% of the 

genera making up the acute copper toxicity database (USEPA 2007a). The species sensitive 

distribution (Figure 1-1) contains the genus mean acute value (GMAV) for 27 genera, 

including 15 species of invertebrates, 22 species of fish, and 1 amphibian species. Each 

GMAV was calculated as the as the geometric mean of the species mean acute values 

(SMAV) for the species within that genera and the SMAV was calculated as the geometric 

mean of the individual value. The selected final acute value of 4.67 ug/L exceeds only one 

of 27 GMAVs (Figure 1-1); i.e., the GMAV for that cladoceran (an invertebrate) Daphnia. 
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Figure 1-1 Ranked freshwater genus mean acute values (GMAVs) for copper (USEPA 2007a). 

Effect thresholds and SMCVs will be calculated for the toxicity data presented in the 

following chapters. The lowest LOAEL will also be noted. Where an EC01, EC10, or 

species sensitivity distribution has been reported, that will also be noted. However, the 

calculation of these other measures of toxicity is not included in this review. 

Selenium is an essential micro-nutrient. Adverse effects occur both when selenium 

concentrations are too low and too high in the diet (see Figure 1-2). At low concentrations, 

selenium is an essential micro-nutrient in animals. If the concentration of selenium in the 

diet is not adequate, selenium deficiency may result. The symptoms of selenium deficiency 

include edema, white muscle disease, reduced growth, reduced disease resistance, liver 

necrosis, morbidity, and increased mortality (National Research Council 1980, 2005, Ullrey 

1992). At high concentrations, however, selenium can cause toxic effects, including 

emaciation, hair/feather loss, reduced reproductive success, reduced growth, developmental 

deformities, and mortality (National Research Council 1980, 2005, Ohlendorf 2003). 
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Figure 1-2 Health vs. dietary selenium concentration (adapted from NRC 2005). 

The occurrence of effects in fish from low dietary selenium concentrations is seen in one of 

the toxicity studies examined in this report. In a study on fingerling channel catfish, Gatlin 

and Wilson (1984) were able to show that dietary selenium concentrations below 

approximately 0.3 mg/kg resulted in reduced weight gain. Above that, there was a plateau in 

the response (i.e., the optimal level) in which increased selenium concentrations did not 

have an adverse effect. However, the plateau ends at approximately 10 mg/kg, with 

concentrations above that also resulting in reduced weight gain (Figure 1-3). This has also 

been observed for calcium, magnesium, manganese, molybdenum, phosphorus, potassium, 

sodium, sulfur, and zinc (NRC 2005). 

 

Figure 1-3 Weight gain in fingerling channel catfish as a function of dietary selenium (Gatlin 

and Wilson 1984). 

Thus, screening values for dietary selenium concentrations should be set above the dietary 

requirement. When dietary concentrations are below the dietary requirements, adverse 

effects could theoretically occur from deficiency. Further, when dietary concentrations are 

near the requirements, a proportion of the population may also be adversely affected by 

selenium deficiency. In turn, these effects could lead to the very reduction in fish and bird 
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populations that the selenium TMDL sought to avoid by reducing selenium concentrations. 

The dietary requirements of fish, birds, pigs, and other mammals have been determined by 

the NRC (1993, 1994, 1998) and will be presented in each of the sections evaluating the 

toxicity to the biota of the North Bay. 
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2 INVERTEBRATE TOXICITY REVIEW

Invertebrates are an integral part of many food chains in the North Bay, supporting many 

species of fish and birds, including species of recreational interest as well as threatened and 

endangered species. This suggests that if selenium negatively impacts the aquatic 

invertebrates in the North Bay, the result may be reduced food availability for the fishes and 

birds of the North Bay. Ultimately, this may lead to a reduction in the population sizes of 

those fishes and birds, even in the absence of direct impacts to the fishes and birds 

themselves from selenium. This suggests that the potential impacts of selenium on aquatic 

invertebrates in the North Bay should be part of the TMDL process. 

While there is a large amount of literature on the toxicity of selenium to fishes, birds, and 

mammals, the chronic toxicity literature is very sparse for aquatic invertebrates. Thus, there 

is very little literature that may help the Water Board to determine whether the aquatic 

invertebrates in the North Bay are impacted by selenium. Similarly, the paucity of 

information on the toxicity of selenium to aquatic invertebrates means that are very few 

published sources of selenium screening values; e.g., no “Consensus-Based Sediment 

Quality Guidelines” (MacDonald et al. 2000), no California sediment quality objectives 

(SWRCB 2007), and only an apparent effect threshold in the NOAA SQuiRTs (Buchman 

1999). 

Despite the paucity of selenium toxicity data for aquatic invertebrates, there are a few 

chronic toxicity studies that may be informative to the Water Board. These studies are listed 

and briefly described below: 

! Alaimo et al. (1994) 

! Bielmeyer et al. (2005) 

! Brasher and Ogle (1993) 

! Debruyn and Chapman (2007) 

! Hyne et al. (2002) 

! Ingersoll et al. (1990) 

! Jensen et al (2007) 

! Malchow et al (1995) 

! Naddy et al (1995) 

! NOAA SQuiRTs (Buchman 1999) 

Alaimo et al. (1994) 

The authors performed chronic toxicity in a laboratory on the larvae of the midge 

Chironomus decorus. Chironomids are a major component of fish and waterfowl diet. 
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Midge eggs were placed in “moderately hard” water and, after hatching, were fed with 

powdered plants that were collected from evaporation ponds in California with varying 

concentrations of selenium. The authors found that increased selenium concentrations in the 

powdered food resulted in lower body weights of the growing midges with a NOAEL and 

LOAEL of approximately 0.5 and 1.6 mg/kg-dw, respectively, in the powdered food (effect 

threshold of 0.9 mg/kg-dw); the NOAEL and LOAEL for selenium in midge tissues were 

both below the detection limit of approximately 0.5 mg/kg-dw. In this experiment, selenium 

was only added via the food, but that also led to increases in selenium in the water column 

over time. 

Bielmeyer et al. (2005) 

The authors performed chronic toxicity tests in a laboratory on the larvae of sea urchins. 

Although the species tested, Diadema antillarum, does not occur in the North Bay, the 

authors state that “sea urchin larvae have been shown to be highly sensitive to metals and 

therefore have been extensively used in marine pollution bioassays” (Bielmeyer et al. 2005). 

Therefore, their results should be applicable to the North Bay in a general sense. The authors 

examined the development of larval sea urchins by exposing them to 0, 5, 10, 20, 36, 73 

!g/L selenium as Na2SeO4 for 40 hours in sea water. The authors found that selenium cause 

abnormal development with an EC50 of 26 !g/L. However, there were no apparent adverse 

effects at up to 20 !g/l (i.e., the NOAEL was 20 !g/L). The experiments design that the 

authors used did not expose the parents to selenium, but only the larvae. In the North Bay, 

both the parents and the offspring will be exposed, which may potentially increase selenium 

sensitivity. 

Brasher and Ogle (1993) 

The authors performed chronic toxicity tests in a laboratory on the freshwater amphipod 

Hyallela azteca. Separate experiments were conducted using sodium selenite at 0, 50, 100, 

200, 300, and 400 !g/L and sodium selenate at 0, 100, 250, 350, 500, and 700 !g/L for 24 

days. Reproduction was reduced at 200 !g/L selenite but no effects were observed for 

selenate.

Debruyn and Chapman (2007) 

The authors compiled existing toxicity data on major effects (i.e., toxic effects that are likely 

to affect a species at the population level, including reproduction and growth) that included 

water, dietary, or tissue selenium concentrations associated with toxic effects to 29 

macroinvertebrate species. Studies reporting dietary and tissue concentrations were pooled 

across freshwater benthic, freshwater planktonic, and terrestrial invertebrate groups. For 

chronic waterborne exposures, the authors found that the amphipod Hyallela was the most 

sensitive with a median lethality reported at 100 !g/L selenite, 200 !g/L selenate, and a 

>50% reduction in reproduction at 100 !g Se/L selenate. Sublethal effects from dietary 

exposures were found to range from 1 to 80 mg/kg-dw. Sublethal effects for tissue 

concentrations of selenium were found to range from 1 to 30 mg/kg-dw. Although the 

authors used the data collected to construct a percentile based species sensitivity distribution 

(not a probability), they did not use the distributions to calculate values protective of a 

percentage of the species observed. 
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Hyne et al. (2002) 

The authors performed chronic toxicity tests in a laboratory on juveniles and adults of the 

estuarine amphipod Corophium sp. Amphipods exposed to sediments spiked with seleno-Lmethionine to give final concentrations of 0, 0.3, 1, 3, and 10 mg/kg-dw. Selenium was not 

added to either the water or the food given to the amphipods and all experiments were 

conducted using seawater. The only response measured in the amphipods was mortality. The 

results of these experiments are summarized in the Table 2-1 below (from Hyne et al. 2002). 

Table 2-1 

Summary of Results from Hyne et al. (2002) 

Lifestage 

Test conditions of 

overlying water 

10-day LC50 sediment 

(dry wt) μg Se/g 

(95% CL) 

10-day NOAEL 

sediment μg Se/g 

 (dry wt) 

10-day LOAEL 

sediment μg Se/g 

(dry wt) 

Juvenile No renewal 1.6 

(1.3-1.9) 

0.84 2.0 

Juvenile Daily renewal 6.3 

(5.5-7.2) 

1.7 5.0 

Adult No renewal 7.6 

(6.8-8.4) 

4.6 11.0 

These results indicate that significant mortality was observed at concentrations as low as 

approximately 2 mg/kg-dw. An effect threshold was calculated for juvenile mortality in 

waters without renewal of 1.3 mg/kg-dw. However, it should be noted that since the authors 

only measured mortality, sublethal effects may be expected at lower sediment 

concentrations. 

Ingersoll et al. (1990) 

The authors performed chronic toxicity tests in a laboratory on the freshwater arthropods 

Daphnia magna and Chironomus riparius. Chironomids are a major component of fish and 

waterfowl diet. The experiments produced LOAELs for inorganic selenium in waterborne 

exposures of 85 !g/L for Daphnia and 837 !g/L for Chironomus. 

Jensen et al. (2007) 

The authors performed freshwater chronic toxicity tests in a laboratory on the larvae of the 

southern house mosquito Culex quinquefasciatus. Culex spp. are algal and bacterial feeders 

and are important as food for many organisms. The larvae were exposed to 2, 4, 8, 16, and 

32 mg/L selenate. The experiments produced a LOAEL for reduced growth of 2 mg/L. 

Malchow et al. (1995) 

The authors performed freshwater acute toxicity tests in a laboratory on the larvae of the 

midge Chironomus decorus. Chironomids are a major component of fish and waterfowl diet. 

The experiments consisted of feeding the midge larvae with a diet of algae (Selenastrum 

capricornutum) cultured in three concentrations of selenite (0, 10, and 40 !g/L) or four 

concentrations of selenate (0, 4, 10, and 40 !g/L). The experiments produced a NOAEL and 

LOAEL for reduced body weight of 1.10 and 2.11 mg/kg-dw, respectively, in algae (effect 

threshold of 1.5 mg/kg-dw); or a NOAEL and LOAEL of <0.25 and 2.55 mg/kg-dw, 

respectively, in the midges (effect threshold of 0.8 mg/kg-dw, assuming the NOAEL was 

0.25). Although this was an acute toxicity test, this is one of the few studies where 

invertebrates were exposed to selenium in the diet. Chronic tests may result in toxicity 

observed at lower concentrations. 
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Naddy et al (1995) 

The authors performed freshwater chronic (9 day) toxicity tests in a laboratory using 

Ceriodaphnia dubia. The daphnia were exposed to <170, 510, 610, and 870 !g/L selenate. 

No effects on survival or reproduction were observed at any concentration. 

NOAA SQuiRTs (Buchman 1999) 

The National Oceanic and Atmospheric Administration has published a set of screening 

values called the Screening Quick Reference Tables (SQuiRTs). The SQuiRTs contain a 

marine sediment apparent effects threshold (AET) protective of amphipods of 1 mg/kg-dw. 

AETs are described as “...the concentration above which adverse biological impacts would 

always be expected... Conversely, adverse impacts are known to occur at levels below the 

AET... AET values were developed for use in Puget Sound (Washington)...” 

2.1 DISCUSSION

“Selenium toxicity is greater in laboratory tests in which organisms are fed than in water 

only tests, because food accumulates Se from the aqueous phase, adding substantially to the 

test organism’s total exposure” (Debruyn and Chapman 2007). Therefore, the results of 

water only tests can result in apparently high toxicity thresholds for invertebrates; e.g., 26 

!g/L for sea urchins (Bielmeyer et al. 2005) and 2,000 !g/L for Culex quinquefasciatus 

(Jensen et al. 2007). Similarly, for sediment dwelling invertebrates, exposures to selenium in 

sediments appears to result in adverse effects at relatively low levels. 

From the few studies presented above that have been performed using either dietary or 

sediment exposures, the following Table 2-2 was constructed: 

Table 2-2 

Effect thresholds for invertebrates exposed to dietary or sediment-bound selenium 

 Exposure Effect Threshold (mg/kg-dw) 

Organism route Diet Sediment Organism Reference 

midge diet 0.9 - <0.5 Alaimo et al. (1995) 

amphipod sediment - 1.3 - Hyne et al. (2002) 

midge diet 1.5 - 0.8 Malchow et al. (1995) 

These data indicate that invertebrates may be affected by relatively low selenium 

concentrations, when exposed to selenium in dietary items or in sediments (for sediment 

dwelling biota). 

Nutritional requirements 

Selenium is a micro-nutrient (see Section 1). When animals do not have enough selenium in 

their diet, selenium deficiency results (NRC 1993, 1994). While the NRC (1993, 1994, 

1998) has not established selenium nutritional requirements for invertebrates, it has 

established nutritional requirements for animals that potentially feed on invertebrates: 0.25-

0.3 mg/kg for fish (NRC 1993), 0.17-0.33 for pigs (NRC 1998), and 0.16-0.22 mg/kg for 

birds (NRC 1994). 

2.2 CONCLUSIONS

The effect thresholds shown in the table above are similar to the few published lower bound 

values available; i.e., the AET for selenium in marine sediments of 1 mg/kg is very close to 
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the effect threshold of 1.3 mg/kg-dw observed by Hyne et al. (2002); and the lower bound 

dietary and organism selenium concentrations published by Debruyn and Chapman (2007) 

of 1 mg/kg-dw each are close to the effect thresholds reported for midges above (0.9-1.5 

mg/kg-dw for diet and <0.5-0.8 mg/kg-dw for organism). All of these values are above the 

dietary requirements for fish, birds, and pigs (NRC 1993, 1994, 1998), although some are 

close. 

Based upon the data presented here, there are several options for screening values protective 

of invertebrates. Those options are as follows (all in units of mg/kg-dw): 

1. Dietary items 

a. Lowest effect threshold: 0.9 

b. Lowest LOAEL: 1.6 

2. Sediment 

a. Lowest effect threshold: 1.3 

b. Lowest LOAEL: 2.0 

c. AET: 1.0 

3. Organism (excluding non-detects) 

a. Lowest effect threshold: 0.8 

b. Lowest LOAEL: 2.55 

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3 FISH TOXICITY REVIEW

The North Bay supports a diverse fish biota. The fish supported include both sportfish and 

threatened and endangered fish species. The five most common sportfish in the North Bay 

are (SFEI 2000; listed in order catch frequency): 

1. Striped bass (Morone saxatilis) 

2. Halibut (Paralichthys californicus) 

3. Jacksmelt (Atherinopsis californiensis) 

4. White sturgeon (Acipenser transmontanus) 

5. White croaker (Genyonemus lineatus) 

In addition to the sportfish listed above, the North Bay supports the following threatened and 

endangered fishes (USFWS 2007): 

1. Chinook salmon (Oncorhynchus tshawytscha) 

2. Delta smelt (Hypomesus transpacificus) 

3. Green sturgeon (Acipenser medirostris) 

4. Longfin smelt (Spirinchus thaleichthys) 

5. Sacramento perch (Archoplites interruptus) 

6. Sacramento splittail (Pogonichthys macrolepidotus) 

7. Steelhead trout (Oncorhynchus mykiss) 

8. Tidewater goby (Eucyclogobius newberryi) 

Selenium in the North Bay may adversely impact these fish by causing reduced production 

of viable eggs, post-hatch mortality, deformities in growing larvae, and various pathological 

effects in the kidneys, liver, heart, and ovaries (Hamilton 2003, 2004; Lemly 2002a). These 

identified effects may lead to population declines of both the sportfish and the threatened 

and endangered fishes in the North Bay. Therefore, as part of the selenium TMDL for the 

North Bay, selenium screening values will be developed that are protective of the species 

listed above. 

As stated in the Introduction (see Section 1), the potential screening values that are proposed 

here are tissue concentrations. For fish, measuring the concentration of selenium in dietary 

items poses a logistical challenge. Mainly, the challenge lies in that many fish feed on a 

wide array of small prey items. Collecting the right kind of prey items, and a sufficient 

number for analysis, can be very time consuming. Instead, it is generally recommended that 

selenium be measured in the fish of concern (Hamilton 2002, 2003, 2004, USEPA 2004). 

Additionally, it is recommended that these measurements be made in whole fish due to 

“...practical reasons of sampling and because a sufficient data base containing chronic 

effects based on whole-body tissue is present in the literature” (USEPA 2004). The amount 

of data for chronic effects based on other tissue types (e.g., ovary, liver, kidney , and 

muscle) is not as extensive and, therefore, not amenable to deriving screening values. 

Further, although “ovaries may be the best tissue to link selenium to reproductive effects... 
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ovarian tissue is also only available seasonally and sometimes difficult to extract in 

quantities sufficient for analysis...” (USEPA 2004). Therefore, whole body concentrations 

in fish are recommended for use as screening values. 

3.1 METHODS

Selenium toxicity data for the fishes that occur in the North Bay would be the most relevant 

for developing selenium fish tissue screening values for the North Bay. However, for the 

fishes in the North Bay, there is very limited toxicity data and almost no toxicity data for 

fishes at all in the salinity range that occurs in the North Bay; i.e., 1 to 33.5 ppt. Therefore, 

this review is largely limited to the evaluation of the freshwater toxicity literature. 

Selenium toxicity studies on fishes were gathered from the scientific literature, as well as 

reports and other “grey literature.” The studies collected were evaluated to select those that 

provided fish tissue concentrations at which toxic effects from chronic selenium exposures 

were observed. To ensure that screening values were not influenced by other variables (e.g., 

other contaminants), only studies with controlled experimental designs were selected. 

The criteria used to identify studies with usable toxicity information are as follows: 

A. provide a NOAEL and/or a LOAEL 

B. chronic 

C. exposed to selenium only 

D. only treatment was selenium exposure level 

E. contained tissue data 

F. fish (and their food) were raised in the laboratory under controlled conditions 

G. used a dietary exposure scenario; mesocosm experiments in which selenium was 

added to the water column meet this requirement 

All studies that were reviewed are listed in Table 3-1. Studies that were determined to be 

unsuitable are noted in Table 3-1 using the designations from the list above. 
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Table 3-1 

Selenium toxicity studies evaluated in this review. Studies that were excluded from review are 

indicated (see text above for an explanation of the criteria). 

Study Reference 

Evaluation

criteria 

failed Study Reference 

Evaluation

criteria 

failed 

1 Bennett et al. (1986) A 41 Hilton and Hodson (1983) 

2 Bertram and Brooks (1986) A 42 Hodson et al. (1980) E 

3 Brandão et al. (1992) A 43 Hodson et al. (1986) A 

4 Chapman (1992) B 44 Holm et al. (2005) F 

5 Cleveland et al. (1993) 45 Hunn et al (1987) G 

6 Coughlan and Velte (1989) F 46 Kennedy et al. (2000) F 

7 Coyle et al. (1993) 47 Kimball (1978) E 

8 de Rosemond et al. (2005) A 48 Klaverkamp et al. (1983) E 

9 Dobbs et al. (1996) 49 Kleinow (1984) A 

10 Doroshov et al. (1992) 50 Kleinow and Brooks (1986a) A 

11 Finger and Bulak (1988) C,D 51 Kleinow and Brooks (1986b) A 

12 Finley (1985) F 52 Lawler et al. (1981) E 

13 Gatlin and Wilson (1984) 53 Lemly (1993a) 

14 Gillespie and Baumann (1986) F 54 Lemly (1993b) F 

15 Gissel-Nielsen and Gissel-Nielsen (1978) E 55 Linville (2006) 

16 Goettl and Davies (1977) E 56 Mehrle et al. (1982) C,D 

17 Goettl and Davies (1978) E 57 Muscatello et al. (2006) F 

18 Hall et al. (1984) C 58 Ogle and Knight (1989) 

19 Hall et al. (1985) B,C,D 59 Pyron and Beitinger (1989) A 

20 Hall et al. (1987) B,C,D 60 Saiki et al. (1992) E 

21 Halter et al. (1980) B 61 Saiki et al. (2004) C 

22 Hamilton et al. (1986) C 62 Schultz and Hermanutz (1990) 

23 Hamilton et al. (1990) 63 Schlenk et al. (2003) B 

24 Hamilton and Wiedmeyer (1990) C 64 Sorensen and Bauer (1983) E 

25 Hamilton et al. (2000) C 65 Sorensen and Bauer (1984) A 

26 Hamilton et al. (2002a) C 66 Sorensen et al. (1982a) A,E 

27 Hamilton et al. (2002b) C,D 67 Sorensen et al. (1982b) A 

28 Hamilton et al. (2005a) C,D 68 Sorensen et al. (1983) A 

29 Hamilton et al. (2005b) C 69 Sorensen et al. (1984) F 

30 Hamilton et al. (2005c) C 70 Sorensen (1988) F 

31 Hamilton et al. (2005d) C 71 Tashjian et al. (2006) 

32 Harrison et al. (1990) A 72 Tashjian et al. (2007) E 

33 Hardy (2003) F 73 Teh et al. (2002) B,E 

34 Hartwell et al. (1987a) A 74 Teh et al. (2004) 

35 Hartwell et al. (1987b) A 75 Vidal et al (2005) 

36 Hermanutz (1992) E 76 Watenpaugh and Beitinger (1985a) A 

37 Hermanutz et al. (1992) 77 Watenpaugh and Beitinger (1985b) B 

38 Hicks et al (1984) 78 Watenpaugh and Beitinger (1985c) A 

39 Hilton et al. (1980) 79 Woock et al. (1987) E 

40 Hilton et al. (1982) A 80 Wise et al. (1993) E 

The reported effects from each study were grouped into one of two categories: major and 

minor effects. Major effects are those that have the potential to impact fish at the organism 

and/or population level (e.g., increased mortality, reduced fecundity, reduced growth, etc.). 

Minor effects are those that are measurable, but are unlikely to result in population level 

effects and/or the long-term impacts on individual fish performance was unclear (e.g., a 

change in an enzyme concentration). Where both major and minor effects were observed at 

the same concentration, the effects was assumed to be major. Only major effects were 

considered in development of screening values. 
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Effect thresholds were calculated as the geometric mean of the NOAEL and LOAEL 

reported for the same effect in an individual study, as is recommended in USEPA water 

quality criteria development guidance (USEPA 1985). Additionally, species mean chronic 

values (SMCVs) were calculated as the geometric mean of the effect thresholds (USEPA 

1985). Studies that did not report a NOAEL were excluded from the calculation of an effect 

threshold and the SMCV. Note that excluding studies in which a NOAEL was not reported 

from the calculation of the SMCV results in lower SMCVs. USEPA (1985) provides the 

following rationale for the use of geometric means to calculate SMCVs: 

“Geometric means rather than arithmetic means are used here because the distributions of 

individual organisms’ sensitivities in toxicity tests on most materials, ... are more likely to 

be lognormal than normal.” 

The calculated SMCVs can then be used to evaluate the relative sensitivity of each species 

of fish to selenium and help determine the appropriate threshold for use. 

In studies where selenium was not measured in whole fish (e.g., selenium was measured in 

liver or muscle), measured concentrations were converted to whole body concentrations 

using the equations presented by USEPA (2004). Additionally, tissue values based upon 

wet-weight values were converted to dry-weight values using either the %moisture values 

presented in the same study or default values from USEPA (2004). All conversion methods 

are described in detail in Appendix A. 

3.2 RESULTS

The toxicity studies from Table 3-1 that were included in the evaluations below included 18 

studies on the following fish species: 

! Bluegill 

! Channel catfish 

! Chinook salmon 

! Fathead minnows 

! Rainbow trout 

! Sacramento splittail 

! White sturgeon 

Bluegill 

Five suitable studies were found on bluegills. Whole body effect thresholds ranged from 3.0 

to 13.8 mg/kg-dw (Table 3-3). 

Channel catfish 

Two suitable studies were identified; i.e., Doroshov et al. (1992) and Gatlin and Wilson 

(1984). In the study by Doroshov et al. (1992), the exposure route was intramuscular 
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injection of selenomethionine. The degree of correlation between intramuscular injection 

and dietary exposures for selenium is unknown. In the study by Gatlin and Wilson (1984), 

selenium exposure was dietary. From these studies, we calculated whole body effect 

thresholds of 2.7 and 3.3 mg/kg-dw, respectively (Table 3-3). 

Chinook salmon 

A single suitable study was identified (Hamilton et al. 1990). That study evaluated selenium 

toxicity in both brackish and freshwaters, the whole body effect thresholds calculated from 

this study were 7.6 mg/kg-dw for freshwater and 17.1 mg/kg-dw for brackish water (Table 

3-3) for fishes fed SeMet. These results indicate that selenium may be less toxic to salmon in 

brackish than in fresh waters. 

Beckon (2007) also analyzed the data from Hamilton et al. (1990) but included the data from 

salmon that were fed field-collected mosquitofish. In our analyses, we excluded the data 

from salmon fed mosquitofish caught in the field as this exposed the salmon to multiple 

contaminants simultaneously which might have negatively influenced survival. Instead of 

calculating an effect threshold from the LOAEL and NOAEL from Hamilton et al. (1990), 

Beckon (2007) performed a regression of selenium concentration vs. survival and estimated 

an EC10 and EC20 of 1.84 and 2.5 mg/kg-dw, respectively. The analysis in shown in the 

figure below. 

Figure 3-1 Biphasic regression analysis of the data from Hamilton et al. (1990) 

It should be noted that performing a regression on the treatment means, instead of the raw 

data from individual test organisms, masks the true variation associated with the 

experimental data; i.e., the confidence limits associated with the regression line in Figure 3-
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1 could be much larger. Further, the nature of the relationship between fish survival and 

whole-body selenium concentrations (i.e., the EC10) may change when using the raw data. 

To examine the effects of using the data for the individual test organisms in the analysis, we 

used the treatment means and standard errors for the same 11 treatments that Beckon (2007) 

used to create the regression analysis in Figure 3-1 and repeated the experiment 10 times in 

a Monte Carlo simulation analysis. Assuming a normal distribution, we generated 2 data 

points per treatment (i.e., the number of data points collected by Hamilton et al. (1990)) and 

performed the same regression analysis. An example graph from one of the simulation 

experiments is provided in Figure 3-2. 

Figure 3-2 Biphasic regression based on simulated data from Hamilton et al. (1990) 

From this simulated data set (Figure 3-2), it can be seen that performing a regression 

analysis on the (simulated) raw data (vs. the treatment means) leads to wider confidence 

intervals on the regression line. For example, the confidence limits on the regression line at 

the lowest treatment level are approximately 48-97% survival, as opposed to the 

approximately 57-80% survival that was calculated using only the treatment means. 

The EC10 and EC20 derived from all 10 simulated experiments are presented in Table 3-2. 

The graphically derived confidence limits are also presented. 
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Table 3-2 

The EC10 and EC20 from 10 simulated experiments and 95% confidence limits (CLs) on salmon 

survival at the EC10 and EC20 

EC10 EC20 

(mg/kgdw) LCL for salmon 

survival at EC10a UCL for salmon 

survival at EC10a (mg/kgdw) 

LCL for salmon 

survival at EC20a UCL for salmon survival at EC20a

1.91 52 83 2.47 47 73 

2.03 55 88 2.61 49 79 

1.59 57 76 2.38 50 68 

1.82 51 74 2.50 46 70 

1.60 47 72 2.26 41 64 

1.88 46 71 2.52 41 64 

1.82 50 77 2.46 45 68 

1.94 48 75 2.74 43 67 

2.42 55 72 3.33 49 64 

1.85 49 74 2.65 44 66 

Notes: 

a - Confidence limits are graphically derived 

UCL – upper confidence limit 

LCL – lower confidence limit 

Further, from Figure 3-2 and Table 3-2, it appears that the EC10 and the EC20 derived by 

Beckon (2007) may not significantly different from each other and that both may not 

significantly different from the no effect level in the regression; i.e., 70% survival. Testing 

for significant differences among the no effect level, the EC10, and the EC20 requires the 

ability to set 95% confidence limits. For non-linear regression analyses, the likelihood-ratio 

method is recommended for calculating confidence limits (see Moerbeek et al. 2004, Sand et 

al. 2006). However, we have not yet taken this next step to calculate confidence limits on the 

EC10 and EC20. 

Fathead minnows 

Three suitable studies were identified for fathead minnows. Whole body effect thresholds 

ranged from 6.0 to 73 mg/kg-dw (Table 3-3). 

Rainbow trout 

Four suitable studies were identified for rainbow trout (Table 3-3). Whole body effect 

thresholds for selenium ranged from 1.7 to 53.1 mg/kg-dw (Table 3-3). Using the treatment 

means from one of the suitable studies (i.e., Hilton et al. 1980; see study 38 in Table 3-3) 

Beckon (2007) derived an EC20 for juvenile rainbow trout of 2.15 mg/kg-dw based on a 

20% reduction in body weight. However, this approach is subject to the same limitations 

discussed above for chinook salmon. 

It should also be noted that Beckon (2007) used the raw data from a study that did not meet 

our selection criteria (i.e., Holm et al. 2005) to derive an EC20 of 2.93 mg/kg-dw for adult 

rainbow trout, based on a 20% reduction in the survival of fry from tout exposed to 

selenium. 

Sacramento splittail 

A single suitable study was identified (Teh et al. 2004). In that study, an effect threshold of 

12.3 mg/kg-dw (muscle) was determined. Using the equations in USEPA (2004) to convert 
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muscle concentrations to whole body concentrations, an effect threshold of 10.8 mg/kg-dw 

(whole body) was calculated (Table 3-3). 

White sturgeon 

Two suitable studies were identified which evaluated the toxicity of selenium to white 

sturgeon (i.e., Linville 2006, Tashjian et al. 2006). Linville (2006) performed two sets of 

experiments: 1) injecting selenium into sturgeon eggs and 2) exposing adult sturgeon to 

dietary selenium. Only the results from the dietary exposure experiment are used in the 

evaluations presented here. We calculated whole body effect thresholds of 6.2 mg/kg-dw 

from Linville (2006) and 18.2 mg/kg-dw from Tashjian et al. (2006) (Table 3-3). William 

Beckon (2008) of the USFWS analyzed the data in Linville (2006) and derived an EC10 for 

selenium in muscle tissue which he then converted to a whole body concentration of 7.50 

mg/kg-dw using a linear regression relating the two tissue types. The regression that 

Beckon (2008) used to convert selenium muscle concentrations to whole body 

concentrations was derived from the treatment means in Tashjian et al. (2006). Tetra Tech 

obtained the raw data from Tashjian et al. (2006) and used that to derive a linear regression 

relating muscle to whole body concentrations (see Appendix A). Using that regression, 

Tetra Tech converted the muscle tissue EC10 derived by Beckon (2008) into a whole body 

concentration of 11.9 mg/kg-dw. However, it should be noted that there were few data 

points available for the calculation of the EC10, making this value somewhat uncertain. The 

regression derived from the raw data from Tashjian et al. (2006) was also used to convert 

Linville’s (2006) muscle NOAEL and LOAEL into whole body concentrations (Table 3-3). 

 
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Table 3-3 

Summary of selenium toxicity studies evaluated in this review. 

Liver1 Muscle1 Whole Body 

Effect Threshold 

(mg/kg-dw) 

Fish Study 

Water 

Type Life stage 

NOAEL 

(mg/kg-dw) 

LOAEL 

(mg/kg-dw) 

NOAEL 

(mg/kg-dw) 

LOAEL 

(mg/kg-dw) 

NOAEL 

(mg/kg-dw) 

LOAEL 

(mg/kg-dw) Liver Muscle Whole Body Endpoint 

bluegill 5 fresh Juvenile 16.8 24.5 5.1 8.4 4.7 7.6 20.3 6.5 6.0 BW, L 

bluegill 7 fresh Adult 30.1 49.1 11.1 21.5 10.0 19.0 38.4 15.4 13.8 R 

bluegill 10 fresh Adult 12.3 26.0 5.8 10.4 5.5 9.3 17.9 7.8 7.1 D 

bluegill 37 fresh Adult 6.0 37.2 1.2 7.2 1.3 6.7 14.9 2.9 3.0 BW, R, S 

bluegill 53 fresh Juvenile 6.3 19.9 1.4 6.3 1.3 7.8 11.2 2.9 3.2 S, BW, L 

channel catfish 10 fresh Adult 7.3 12.5 2.1 3.8 2.0 3.6 9.6 2.8 2.7 S 

channel catfish 13 fresh Fingerling - - 3.0 3.5 3.1 3.5 - 3.3 3.3 BW 

chinook salmon 23 fresh Larvae2 18.8 31.9 5.9 12.0 5.4 10.8 24.5 8.4 7.6 BW, L 

chinook salmon 23 brackish Juvenile2 35.9 57.2 14.1 26.4 12.6 23.2 45.3 19.3 17.1 BW, L, S 

fathead minnow 9 fresh Larvae - 121.0 - 72.8 - 62.0 - - - BW 

fathead minnow 58 fresh Juvenile 18.8 21.9 5.7 6.6 5.4 6.6 20.3 6.1 6.0 BW 

fathead minnow 62 fresh Egg - - - - 1.6 19.6 - - 5.5 D 

rainbow trout 75 fresh Larvae 6.1 9.9 1.3 2.5 1.2 2.3 7.8 1.8 1.7 BW, L 

rainbow trout 38 fresh Juvenile 153.2 197.2 - - 46.8 60.3 173.8 - 53.1 BW, FG, S 

rainbow trout 39 fresh Juvenile 42.0 95.0 - - 12.6 28.9 63.2 - 19.1 BW, S, FG 

rainbow trout 41 fresh Juvenile3 21.0 71.7 1.0 4.3 1.0 4.0 38.8 2.1 2.0 BW, FG 

Sacramento splittail 74 fresh Juvenile 23.0 26.8 10.1 15.1 9.0 12.9 24.8 12.3 10.8 D 

white sturgeon 55 fresh Adult 1.4 10.4 1.3 12.1 3.1 12.3 3.9 4.0 6.2 D 

white sturgeon 71 fresh Juvenile 22.0 37.4 22.9 36.8 14.7 22.5 28.7 29.0 18.2 BW, SA 

Notes: 

1 - shaded cells represent concentrations predicted from measurements in other tissues (see Appendix A) 

2 - only results from the SeMet dietary exposure part of the study are used here 

3 - data from the low carbohydrate diet were not used as the authors reported that reduced weight of fish on this diet was likely due to food avoidance, not toxicity 

BW = Body Weight 

D = Deformities 

FG = Feed:Gain ratio 

L = Length 

R= Reproduction 

S = Survival 

SA = Swimming Activity 
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Using the data presented in Table 3-3, SMCVs were calculated for each fish species. For 

species where only one toxicity study was reported, the SMCV is equal to the effect 

threshold from that study. The SMCVs are shown in the Table 3-4. 

Table 3-4 

Species mean chronic values calculated from the toxicity studies evaluated above. 

SMCV (mg/kg-dw) 

Fish Species Water Type Liver Muscle Whole body 

Bluegill Lepomis macrochirus fresh 18.8 5.8 5.6 

Channel catfish Ictalurus punctatus fresh 9.6 3.0 3.0 

Chinook salmon Oncorhynchus tshawytscha fresh 24.5 8.4 7.6 

Chinook salmon Oncorhynchus tshawytscha brackish 45.3 19.3 17.1 

Fathead minnow Pimephales promelas fresh 20.3 6.1 6.0 

Rainbow trout Oncorhynchus mykiss fresh 42.7 1.9 7.6 

Sacramento splittail Pogonichthys macrolepidotus fresh 24.8 12.3 10.8 

White sturgeon Acipenser transmontanus fresh 10.5 10.8 10.6 

Note: shaded cells are based (at least partly) on regressions, not measurements 

Life History differences 

To determine whether there is a trend in the sensitivity to selenium over the lifespan of a 

fish, the data gathered were segregated by life-history stage (e.g., adult, juvenile/fingerling, 

fry/larvae, egg). There were little data for fish larvae/eggs, and the majority of the available 

data are for effects on juveniles and adults. From the available data (Figures 3-3, 3-4) there 

does not appear to be a relationship between age and selenium sensitivity. A relationship 

may be present, but there is insufficient data available to adequately evaluate the existence 

of a pattern. 
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Centrarchids

Juvenile Adult

Stage

2

4

6

8

10

12

14

16

Selenium (mg/kg dw) in Centrarchids (whole body)

 Major effects

Figure 3-3 Effect thresholds for centrarchids (i.e., bluegill, green sunfish, and redear 

sunfish) from the selected studies grouped by life history stage. 

Fry/larvae Juvenile

Stage

10

20

30

40

50

60

70

Selenium (mg/kg dw) in Rainbow Trout (whole body)

 Minor effects

 Major effects

Figure 3-4 Effect thresholds for rainbow trout from the selected studies grouped by life 

history stage. 
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Since there were too few suitable studies to evaluate the potential effect of life history stage 

on selenium toxicity to fish, some of the criteria that had been used to exclude studies were 

relaxed. Specifically, criteria F and G of Table 3-1 were relaxed to increase our ability to 

evaluate life history stage. However, this means that studies that were included where fishes 

were either fed food that was not reared under controlled conditions or where fishes were 

exposed to selenium in water only. The results are shown in Figures 3-5 and 3-6. Note that 

the studies that were added by relaxing the selection criteria were not included in the 

calculation of SMCVs or screening values. 

egg Fry/larvae Juvenile Adult

Stage

0

5

10

15

20

25

30

35

Selenium (mg/kg dw) in Centrarchids (whole body)

 Major effects

Figure 3-5 Effect thresholds for centrarchids after relaxing the selection criteria. 
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Egg Fry/larvae Juvenile Adult

Stage

0.4

0.5

0.6

0.7

0.8 0.9 1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0 9.0 10.0

20.0

30.0

40.0

50.0

60.0

70.0

Selenium (mg/kg dw) in Rainbow Trout (whole body)

 Minor effects

 Major effects

Figure 3-6 Effect thresholds for rainbow trout after relaxing the selection criteria. 

Although the range of effect thresholds in Figure 3-5 overlap, the mean effect threshold for 

juvenile centrarchids is 1.7 times lower than for adults. No trend was apparent in the data 

available for rainbow trout. Overall, it does not appear that eggs are more sensitive than 

juvenile or larval fish. There are several potential reasons that the available data do not show 

that eggs are more sensitive than juvenile/larval fish: 1) the endpoints measured were not 

sensitive enough, 2) the type of selenium exposure was not conducive to observing effects 

on eggs (e.g., waterborne exposures), or 3) fish eggs may indeed be less sensitive to toxic 

chemicals than other life history stages (Finn 2007). 

3.3 DISCUSSION

80 selenium toxicity studies were reviewed in an effort to derive selenium screening values 

for species representative of fishes in the North Bay (Table 3-1). Based on this review, 

species which are appropriate for use in developing a fish tissue screening values for the 

North Bay, and for which usable selenium toxicity data are available, include: bluegill 

(Lepomis macrochirus), fathead minnow (Pimephales promelas), rainbow trout 

(Oncorhynchus mykiss), Chinook salmon (Oncorhynchus tshawytscha), Sacramento splittail 

(Pogonichthys macrolepidotus), and white sturgeon (Acipenser transmontanus). 

With the exception of the study on the impacts of selenium on Chinook salmon in fresh and 

brackish water (Hamilton et al. 1990), all usable studies performed to date on the effects of 

selenium on fish have been performed in freshwater. In addition to these species, selenium 

toxicity data was sought for striped bass, flounder, and delta smelt. However, no usable 

selenium toxicity data were available for these fish, nor were data available for any other 

closely related species (e.g., pond smelt, Hypomesus olidus). 
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Based on the available data, it appears that the species of concern in the North Bay for 

which toxicity data are available (i.e., Sacramento splittail and white sturgeon) are less 

sensitive to selenium than bluegill (Table 3-4), a species that has been observed to be 

particularly sensitive to selenium in a number of studies (e.g., Lemly 1993a) and was 

selected by USEPA as the basis of the current draft water quality criterion for selenium 

(USEPA 2004). Further, channel catfish appear to be even more sensitive to selenium than 

bluegill. However, with only one dietary and one injection toxicity study available on 

catfish, confidence in the sensitivity of channel catfish to selenium is limited. 

Other selenium screening values 

The selection of an appropriate selenium fish tissue screening value, protective of toxic 

effects in both individual fish and populations, is an area of active debate (Chapman 2007; 

Hamilton 2003, 2004; Lemly and Skorupa 2007; McDonald and Chapman 2007; Skorupa et 

al. 2004; USEPA 2004). Numerous screening values for selenium in fish tissues have been 

proposed in the past. These values are summarized in Table 3-5. 

Table 3-5 

Whole body fish tissue screening levels (mg/kg-dw) published in the scientific literature. 

Reference No effecta Level of concernb Toxicity Thresholdc

Lemly and Smith (1987) 12 

Henderson et al (1995) 4 4-12 12 

Stephens et al. (1997) 2-3 4 4 

USDOI (1998) 2-3 2-4 4 

DeForest et al (1999) 6 (cold water) 

DeForest et al (1999) 9 (warm water) 

Lemly (1993c, 1996, 2002b) 4 

Hamilton (2003) 4 

Ohlendorf (2003) 4 

USEPA (2004) 7.9 

Notes: 

a - Concentrations less than this value produce no discernible adverse effects on fish or wildlife and are typical of 

background concentrations in uncontaminated environments (USDOI 1998). 

b - Concentrations in this range rarely result in discernible adverse effects on some fish or wildlife species (USDOI 

1998). 

c - Concentrations greater than this value may result in adverse effects on some fish or wildlife species (USDOI 

1998). 

The earliest screening values were 12 mg/kg-dw (see Lemly and Smith (1987), Henderson et 

al. (1995)) (Table 3-5). However, as more data became available, most authors generally 

recommended a screening value of 4 mg/kg-dw (Table 3-5). In part, those screening values 

were derived to be protective of the lowest concentration at which selenium exposures 

resulted in toxicity in a single study. Thus, the approach used to derive the screening values 

in the scientific literature has generally been to use the lowest whole body values from Table 

3-3. Further, the approach in the scientific literature has been to derive a screening value 

protective of all fish species across the entire United States and not just the North Bay. 

There have also been several critiques of the USEPA (2004) draft that have noted statistical 

and interpretational errors in the draft that potentially affect the draft screening value (Lemly 

and Skorupa 2007, Skorupa et al. 2004). 
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Presser and Luoma (2006) also reviewed the toxicity literature and concluded that wholebody fish tissue concern levels range from 1.5 to 6 mg/kg-dw. 

Warm water vs. cold water screening values 

One of the major topics of discussion in the recent past has been whether cold water fish are 

more sensitive to selenium that warm water fish. Recent reviews have not found support for 

the cold vs. warm water distinction (e.g., Hamilton 2003, Chapman 2007). 

Effect of sulfate on selenium toxicity 

Acute toxicity tests have shown that sulfate reduces the potential for bioaccumulation and 

toxicity of selenium to algae, aquatic invertebrates, and fish (see Brix et al. 2001, Schlenk et 

al. 2003, USEPA 2004). USEPA (2004) provides the following discussion on the effects of 

sulfate on selenium toxicity: 

“The toxicity of a number of metals (e.g., copper and cadmium) to aquatic organisms is 

related to the concentration of hardness in the water. The toxicity of these metals to many 

different aquatic species has been shown to decrease as the hardness concentration 

increases. A similar relationship also has been recognized between selenate and dissolved 

sulfate in freshwater (a similar relationship is not evident between selenite and sulfate or 

between either form of selenium and hardness). The studies reviewed in this document 

indicate that, as the concentration of sulfate increases, the acute toxicity of selenate is 

reduced (less toxic). Selenate acute toxicity tests conducted at different levels of dissolved 

sulfate are available with C. dubia, D. magna, H. azteca, G. pseudolimnaeus, chinook 

salmon and fathead minnows (Table 1a [of USEPA 2004]). These data indicate that, in 

general, selenate is more toxic to these species in low sulfate water than in higher sulfate 

water.” 

Since sulfate levels should be higher in brackish and marine waters than in freshwaters, 

screening values based on freshwater selenium toxicity studies are likely to be protective of 

potential impacts to fishes in estuarine and marine environments. There are several possible 

reasons why sulfate may reduce the toxicity of selenate, including 1) the ability of sulfate, 

calcium, or magnesium to prevent the interaction of selenomethionine with critical 

subcellular proteins by maintaining the cellular redox potential (Schlenk et al. 2003) and 2) 

direct competition at the cell uptake site as selenate and sulfate are structurally similar group 

VI oxyanions of the form XO4; experiments with bacteria indicate that selenate and sulfate 

have a common membrane carrier and that active transport by this carrier is the only means 

by which selenate may enter a cell (Brix et al. 2001). 

Potential screening values for the North Bay 

The objective of this review of the selenium toxicity information available for fishes is to 

provide a technical basis for the selection of an appropriate selenium fish tissue screening 

value for in the North Bay. Since there are limited selenium toxicity data available for fish 

of concern for the North Bay (e.g., no suitable selenium toxicity data for delta smelt, halibut, 

jacksmelt, striped bass, Tidewater goby, or white croaker), it is not possible to develop 

screening values that take into account the effects of selenium on all fish species of potential 

concern in the North Bay. 
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The SMCVs for species such as Chinook salmon, rainbow trout, Sacramento splittail, and 

white sturgeon indicate that these species are less sensitive to selenium than bluegill and 

channel catfish (Table 3-4), which appear to be fairly sensitive to selenium. Thus, using a 

screening value for the North Bay based on either bluegill (e.g., SMCV of 5.6 mg/kg-dw 

whole body) or channel catfish (e.g., SMCV of 3.0 mg/kg-dw whole body) should be 

protective of most (if not all) fish species, while not setting an unrealistically low threshold 

for potential effects. 

The single study that evaluated the effects of selenium in both brackish and freshwater 

(Hamilton et al. 1990), indicated that Chinook salmon were more than twice as sensitive to 

selenium in freshwater as in brackish water. However, this study examined the effects of 

selenium on larvae in fresh water and juveniles in brackish water. Therefore, the greater 

sensitivity observed in freshwater with larvae could also be due to life history differences in 

selenium sensitivity. 

Table 3-4 also presents SMCVs for selenium in liver and muscle. Liver concentrations 

respond rapidly to changes in the environment and may be useful in detecting short term 

selenium fluctuations. Muscle concentrations may be non-destructively measured in 

relatively large fishes by collecting and analyzing muscle plugs. This may be desirable as it 

reduces the impact of selenium monitoring on fish populations, which is especially 

important for special status species. However, it should be noted that the regressions used 

here to convert selenium concentrations measured in one tissue type to another tissue type 

are not specific to the species and populations considered here, with the exception of white 

sturgeon (see Appendix A). Therefore, using screening values that were calculated using the 

regressions from USEPA (2004) incorporates some uncertainty. 

Beckon (2007) 

In contrast to the approach used here (i.e., calculating effect thresholds and SMCVs), 

Beckon (2007) derived an EC10 and an EC20 (of 1.84 and 2.5 mg/kg-dw, respectively) for 

the survival of salmon fry by performing a regression on the treatment means in Hamilton 

(et al. 1990) (see above). However, it should be noted that performing a regression on the 

treatment means artificially reduces the variation. Further, the shape of the relationship may 

change when using the raw data. Thus, the EC10 and EC20 derived by Beckon (2007) 

should be viewed as approximations. It is also unknown whether the EC10 and EC20 are 

significantly different from each other. 

Nutritional requirements 

Selenium is a micro-nutrient (see Section 1). When animals do not have enough selenium in 

their diet, selenium deficiency results (NRC 1993, 1994). For fish, nutrient requirements are 

given in terms of the concentration in their feed, not in fish tissue, as is being used here to 

derive screening values. Nonetheless, the nutritional requirements of two species of fish 

have been established and are provided here for reference: 0.25 mg/kg for channel catfish 

and 0.3 mg/kg for rainbow trout (NRC 1993). These concentrations are given in an “as fed 

basis.” Fish feed is usually in the form of dry pellets and probably has a moisture content of 

approximately 10%, meaning the dietary requirements listed above in mg/kg-as fed do not 

differ substantially from mg/kg-dw. 
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Additional considerations 

Issues to be considered in the collection and interpretation of samples include: 

! tissue concentrations in fishes may vary with the time of year sampled 

! tissue concentrations may vary by geographic location 

! if sampling dietary items of birds, selenium concentrations may also differ among 

their prey items 

! number of samples to collect 

! the frequency of collections 

! whether samples should be composited 

! how to compare measured concentrations to screening values 

These issues are discussed in greater detail in Section 6. 

The studies used in the determination of the screening value presented above, and all 

screening values obtained from the literature, were not derived using the populations of fish 

that occur in the North Bay nor using the environmental conditions in the North Bay. This 

means that there is some uncertainty as to whether any screening criteria are actually 

applicable. Similarly, McDonald and Chapman (2007), recommend that screening values 

only be used to determine if further evaluation is necessary. They recommend that if fish 

tissue concentrations exceed an appropriate screening value that 1) reproductive toxicity 

testing be conducted for the species of concern and that if reproductive toxicity testing 

indicates that the species has been negatively affected by selenium that 2) fish population 

surveys be conducted. McDonald and Chapman (2007) provide recommendations on how 

both reproductive toxicity testing and fish population surveys can be integrated into a 

selenium monitoring program. 

Future research and data gaps 

Several data gaps and areas where further research may be beneficial to the Water Board 

were identified during the evaluation of selenium toxicity to fish in the North Bay, as 

follows: 

1. Although it is believed that selenium should be less toxic to fish in marine and 

estuarine systems than freshwater systems, the data to show this are lacking. 

Toxicity experiment performed with representative fish species in using multiple 

salinities over the range that occurs in the North Bay would provide valuable 

information. 

2. Toxicity studies have not been performed on most of the species of concern (which 

includes endangered species) in the North Bay. Thus, this report has assumed that 

toxicity studies on bluegill, rainbow trout, chinook salmon, channel catfish, and 
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fathead minnows are representative of the species of concern. The species of 

concern may be more or less sensitive. 

3. Channel catfish appear to be one of the most sensitive species. However, only two 

studies were available. If the Water Board believes that channel catfish are 

representative of the fish in the North Bay, additional toxicity studies using channel 

catfish would be warranted. 

4. The existing toxicity data suggests that younger life history stages are more 

sensitive to selenium, but most selenium toxicity studies have focused on only a 

single life history stage. Thus, it is possible that some species may be more 

sensitive to selenium than shown here. 

5. Species-specific regressions to convert selenium concentrations from one tissue 

type to another are lacking for all of the fish species evaluated here, except white 

sturgeon. Therefore, the regressions from USEPA (2004) for bluegill were used for 

most species. Since the applicability of these regressions to other species is 

unknown, screening values derived using the bluegill tissue relationships should be 

viewed as approximations and may need to be revisited as new data becomes 

available. Further, if the Water Board decides to monitor fish tissue concentrations 

by collecting muscle plugs, it would be beneficial to develop species-specific 

regressions. 

6. A detailed re-analysis of the published toxicity data using, for example, biphasic 

regressions (Beckon 2007, Beckon et al. 2008) would provide a greater level of 

confidence in estimates of the threshold for toxic effects of selenium vs. a LOAEL, 

NOAEL, or effect threshold. However, if only the treatment means are available for 

the analyses, it should be remembered that this type of analysis will mask the true 

variability in the data. 

3.4 CONCLUSIONS

Exposure to elevated levels of selenium can cause adverse effects in fish. Two of the species 

that occur in the North Bay that have been tested (i.e., Sacramento splittail and white 

sturgeon) appear to be less sensitive than other freshwater fish. While it has generally been 

believed that bluegill are one of the more sensitive fish species, the two studies available on 

channel catfish indicates that they may be more sensitive than bluegill. However, one of the 

most important concerns for selenium toxicity to fish remains an open question at the end of 

this review: is selenium less toxic to fish in marine and brackish waters than freshwater? The 

available data supports the assertion that selenium should be less toxic to fish in marine and 

brackish waters, but it is not possible to quantify the difference at this point. 

Numerous selenium screening values protective of fish have been proposed. For whole body 

fish selenium concentrations, the recommended screening values are usually in the range of 

2-4 mg/kg-dw (Table 3-5). Additionally, it has been proposed that separate screening values 

be used for cold water fish vs. warm water fish, although recent reviews have not found 

support this distinction (e.g., Hamilton 2003, Chapman 2007). 

Based upon the data presented here, there are several options for selenium fish tissue 

screening values. Those options are as follows (all in units of mg/kg-dw whole body): 
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1. The lowest SMCV: 3.0 (Table 3-4) 

2. The salmon EC20 from Beckon (2007) of 2.5 

3. Lowest concentration at which adverse effects were reported (Table 3-3) 

a. LOAEL of 2.3 

4. The lowest LOAELs for fish in the North Bay (Table 3-3) 

a. 12.9 for Sacramento splittail 

b. 12.3 for white sturgeon 

5. The lowest effect thresholds for fish in the North Bay (Tables 3-3, 3-4) 

a.10.8 for Sacramento splittail 

b.6.2 for white sturgeon 

6. The lowest SMCVs for fish in the North Bay (Table 3-4) 

a.10.8 for Sacramento splittail 

b.10.6 for white sturgeon 

7. The white sturgeon EC10 calculated by Beckon (2008) from the data in Linville (2006) 

of 11.9 

The choice of using an SMCV, LOAEL, or effect threshold is dependent upon both the level 

of risk and uncertainty the Water Board wishes to accept. This issue was discussed in more 

detail in the Introduction. It is worth noting that of these values, only the LOAELs represent 

concentration at which effects have been demonstrated. However, since only a few 

concentrations have been tested, effects may actually occur at lower concentrations. 

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4 AVIAN TOXICITY REVIEW

The waters of the North Bay supports a diverse avian biota. Some birds are residents in the 

Bay all year round, while many others rely on the North Bay as part of their annual 

migrations. The birds supported include both game birds as well as and threatened and 

endangered species. The North Bay supports the following threatened and endangered bird 

species (USFWS 2007): 

! American bittern (Botaurus lentiginosus) 

! American peregrine falcon (Falco peregrinus anatum) 

! American white pelican (Pelecanus erythrorhynchos) 

! bald eagle (Haliaeetus leucocephalus) 

! black tern (Chlidonias niger) 

! California black rail (Laterallus jamaicensis coturniculus) 

! California brown pelican (Pelecanus occidentalis californicus) 

! California clapper rail (Rallus longirostris obsoletus) 

! California least tern (Sterna antillarum browni) 

! double-crested cormorant (Phalacrocorax auritus) 

! marbled murrelet (Brachyramphus marmoratus) 

! mountain plover (Charadrius montanus) 

! northern harrier (Circus cyaneus) 

! osprey (Pandion haliaetus) 

! prairie falcon (Falco mexicanus) 

! tricolored blackbird (Agelaius tricolor) 

! western least bittern (Ixobrychus exilis hesperis) 

! western snowy plover (Charadrius alexandrinus) 

! white-faced ibis (Plegadis chihi) 

! white-tailed kite (Elanus leucurus) 
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The North Bay supports a number of birds that are not classified as threatened and 

endangered but are covered under the Migratory Bird Treaty Act (MBTA), including 

(USFWS 2007): 

! Aleutian Canada goose (Branta canadensis leucopareia) 

! black scoter (Melanitta nigra) 

! Caspian tern (Sterna caspia) 

! greater scaup (Aythya marila) 

! lesser scaup (Aythya affinis) 

! surf scoter (Melanitta perspicillata) 

! white-winged scoter (Melanitta fusca) 

Lastly, the North Bay supports several species of water fowl that are of recreational interest. 

While the California Department of Fish and Game (DFG) does not provide any information 

on which species are harvested in the North Bay, they do provide limits (i.e., “bag limits”) 

on how many can be harvested at one time (DFG (2007). These limits, which indicate which 

water fowl species are of recreational interest, are provided below: 

! canvasback (Aythya valisineria): 1 

! American coot (Fulica americana): 25 

! Ross’s goose (Chen rossii): 4 

! snow goose (Chen caerulescens): 4 

! white-fronted goose (Anser albifrons): 4 

! cackling goose (Branta hutchinsii): 6 

! Aleutian goose (Branta canadensis leucopareia): 6 

! Canada goose (Branta canadensis): 4 

! mallard (Anas platyrhynchos): 7 but not more than 2 hens 

! common moorhen (Gallinula chloropus): 25 

! pintail (Anas acuta): 1 

! redhead (Aythya americana): 2 
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! scaup (Athya affinis and Athya marila): 3 

! merganser (Mergus spp. and Lophodytes cucullatus): 7 

The USFWS (2007) evaluated the bird species found in the North Bay that are covered 

under the MBTA as well as the threatened and endangered species and determined that the 

following species were most at risk from potential exposures to selenium: 

! bald eagle (Haliaeetus leucocephalus) 

! black scoter (Melanitta nigra) 

! California clapper rail (Rallus longirostris obsoletus) 

! greater scaup (Aythya marila) 

! lesser scaup (Aythya affinis) 

! surf scoter (Melanitta perspicillata) 

! white-winged scoter (Melanitta fusca) 

Among birds, exposures to elevated levels of selenium can induce gross abnormalities 

during development, egg inviability, chick mortality (Adams et al. 2003), reduced weight 

gain, reduced adult condition, and reduced adult survival (NRC 2005, Ohlendorf 2003). 

When the selenium concentrations are high enough in a large enough area, these effects can 

lead to population declines for the most affected species. 

The step in the effort to determine whether birds may be adversely impacted by selenium in 

the North Bay is to evaluate the toxicity of selenium to the marine mammals. Next, potential 

selenium screening values for the species in the North Bay are proposed. Lastly, the 

screening values are compared to nutritional requirements. 

4.1 METHODS

Selenium toxicity data for the bird species that the USFWS (2007) has identified as being 

most at risk from selenium in the North Bay (see list above) would be the most relevant for 

the Water Board’s development of a TMDL. However, selenium toxicity studies have not 

been performed for any of the species identified by the USFWS (2007) as being most at risk 

from selenium in the North Bay. Therefore, in this section of the report, we review the 

available avian selenium toxicity literature. 

For birds, selenium studies have generally focused on 1) determining a toxic concentration 

in the diet, 2) determining a toxic concentration in eggs, or 3) improving bird “performance” 

on a poultry farm. We reviewed papers published in the scientific literature for all three 

types of data. Studies were excluded from the consideration for the selection of a screening 

value (see Table 4-1) if they: 

A. did not provide a LOAEL 
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B. were not chronic 

C. exposed birds to multiple contaminants simultaneously 

D. exposed birds to multiple treatments simultaneously 

E. field studies 

Using the filtering criteria above, only studies where the reported effects can unambiguously 

be attributed to selenium were reviewed. Some of the poultry performance studies were not 

performed using dietary concentrations high enough to cause toxic effects. These studies 

will be used to set a lower bound on screening values. Most of the studies describing the 

early laboratory work on the effects of selenium on poultry species are no longer readily 

accessible. When the original studies are not readily available, the reviews that summarize 

the results of the original studies are cited. 

The reported effects from each study were grouped into one of two categories: major and 

minor effects. Major effects are those that have the potential to impact birds at the organism 

and/or population level (e.g., increased mortality, reduced fecundity, reduced growth, etc.). 

Minor effects are those that are measurable, but are unlikely to result in population level 

effects and/or the long-term impacts on individual bird performance was unclear (e.g., a 

change in an enzyme concentration). Where both major and minor effects were observed at 

the same concentration, the effects was assumed to be major. Only major effects were 

considered in development of screening values. 

All studies that were reviewed are listed in Table 4-1. Studies that were determined to be 

unsuitable are noted in Table 4-1 using the designations from the list above. The 21 suitable 

studies selected are presented in Table 4-2. 
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Table 4-1 

Selenium toxicity studies evaluated in this review. Studies that were excluded from review are 

indicated in the table (see text above for an explanation of the criteria). 

Study Reference 

Evaluation 

criteria failed Study Reference 

Evaluation 

criteria failed 

1 Albers et al. (1996) 24 Latshaw et al. (2004) 

2 Anteau et al. (2007) A,C,E 25 Lowry and Baker (1989) 

3 Biswas et al. (2006) A 26 Ort and Latshaw (1978) 

4 Elzubier and Davis (1988) 27 O'Toole and Raisbeck (1997) 

5 Fairbrother et al. (1994) C,E 28 Pappas et al. (2005) A 

6 Franson et al. (2007) 29 Paton et al. (2002) A 

7 Green and Albers (1997) 30 Payne et al. (2005) A 

8 Harding (2007) E 31 Ratti et al. (2006) A,E 

9 Heinz and Hoffman (1996) 32 Rattner et al (2000) A,C,E 

10 Heinz and Hoffman (1998) 33 Ryu et al. (2005) A 

11 Heinz et al. (1987) 34 Sahin and Kucuk (2001) A 

12 Heinz et al. (1988) 35 Santolo et al. (1999) A 

13 Heinz et al. (1989) 36 Stanley et al. (1994) 

14 Heinz et al. (1996) 37 Stanley et al. (1996) 

15 Hoffman and Heinz (1998) 38 Stoewsand et al. (1978a) A,C 

16 Hoffman et al. (1991) 39 Stoewsand et al. (1978b) A,C 

17 Hoffman et al. (1992a) 40 Surai et al. (2006) A 

18 Hoffman et al. (1992b) 41 Takekawa et al. (2002) A,C,E 

19 Hoffman et al. (1996) 42 Wayland et al. (2002) A,C,E 

20 Hoffman et al. (1998) C,E 43 Wiemeyer and Hoffman (1996) 

21 Hoffman et al. (2002) C,E 44 Wilson et al. (2007) A,C,D,E 

22 Jensen (1975) 45 Yamamoto and Santolo (2000) A 

23 Kinder et al. (1995) D 46 Yamamoto et al. (1998) A 

To be consistent with the approach used for developing screening values for fish, effect 

thresholds were calculated as the geometric mean of the NOAEL and LOAEL reported for 

the same effect in an individual study. Additionally, species mean chronic values (SMCVs) 

were calculated as the geometric mean of the effect thresholds (USEPA 1985). Studies in 

which a zero concentration was reported in the control were assumed to have used 0.2 

mg/kg-dw, as this is the dietary requirement for birds (NRC 1984). Studies that did not 

report a NOAEL were excluded from the calculation of an effect threshold and the SMCV. 

Note that excluding studies in which a NOAEL was not reported from the calculation of the 

SMCV results in lower SMCVs. 

Conversions from dry weight to wet weight are detailed in Appendix A. 

4.2 RESULTS

Toxicity studies 

The toxicity studies from Table 4-1 that were included in the evaluations below included the 

following bird species: 

! Chickens 

! Eiders 

! Kestrels 

! Pheasants 
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! Mallards 

! Screech owls 

In contrast to the studies on fish, where whole body selenium concentrations are usually 

measured, the toxicity studies performed on birds have usually measured the selenium 

content of the feed given to the bird. Therefore, the screening values derived here are for 

dietary selenium concentrations. 

Chickens 

Four studies were identified for domestic chickens (Table 4-2). Only one study (Ort and 

Latshaw 1978) examined the effect of selenium on the most sensitive endpoint; i.e., egg 

hatchability. It is also worth noting that Ort and Latshaw (1978) review two earlier studies 

not available to us that show 1) “egg weight, fertility and hatchability were all significantly 

decreased by 8 ppm [dietary] selenium,” and 2) “Embryonic development was adversely 

affected by 10 ppm dietary selenium but not by 5 ppm.”
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Table 4-2 

Summary of dietary selenium toxicity studies evaluated in this review. 

(mg/kg-dw) Effect 

Bird Study Life stage Dietary form NOAEL LOAEL Effect threshold Major Minor Endpoint 

Chicken 4 Chick selenite 0.15 10 1.2 X BW 

Chicken 22 Chick selenite 0.9 4.3 2.0 X BW 

Chicken 22 Chick selenite 4.3 13.5 7.6 X BW, S 

Chicken 25 Chick selenite; SeMet 0.2 15 1.7 X BW 

Chicken 26 Adult selenite 3 5 3.9 X R 

Common eider 6 Adult SeMet 20.6 57.7 34.5 X BW 

Mallard 1 Adult SeMet 22.3 44.7 31.6 X C 

Mallard 7 Adult SeMet 11.2 22.3 15.8 X S 

Mallard 9 Adult SeMet 0.6 10.6 2.6 X R 

Mallard 10 Adult SeMet 0.4 9.8 2.1 X R 

Mallard 11 Adult selenite 10.9 27.3 17.3 X BW, R 

Mallard 11 Adult SeMet 0.2 10.9 1.5 X R 

Mallard 12 Duckling selenite 10.9 21.6 15.3 X BW 

Mallard 12 Duckling SeMet 10.9 21.6 15.3 X BW 

Mallard 13 Adult SeMet 4.6 9.0 6.4 X R 

Mallard 14 Duckling SeMet 17.0 33.7 24.0 X BW 

Mallard 15 Adult SeMet 0.4 11.6 2.3 X E 

Mallard 16 Duckling SeMet; low 

protein 

0.2 16.9 1.9 X BW 

Mallard 16 Duckling SeMet 16.9 66.9 33.6 X BW, S 

Mallard 17 Duckling SeMet 0.2 16.9 1.9 X BW 

Mallard 18 Duckling SeMet 0.2 16.5 1.9 X E 

Mallard 18 Duckling SeMet 16.5 65.4 32.9 X BW 

Mallard 18 Duckling SeMet; low 

protein 

16.5 65.4 32.9 X S 

Mallard 18 Duckling SeMet; low 

protein 

0.2 16.5 1.9 X E 

Mallard 18 Duckling SeMet; high 

protein 

0.2 16.5 1.9 X BW 

Mallard 19 Duckling SeMet, Se yeast, 

Se in wheat 

0.2 16.9 1.9 X E 

Mallard 27 Flightling SeMet 13.8 33.8 21.6 X BW, S 

Mallard 36 Adult SeMet 0.37 6.5 1.6 X R 

Mallard 37 Adult SeMet 3.9 7.8 5.5 X BW, R 

Pheasant1 24 Adult unknown 0.4 9.3 1.9 X R, S 

Screech owl 43 adult SeMet 8.8 30 16.2 X BW, R 

Screech owl 43 adult SeMet 0.3 8.8 1.6 X E 

Notes: 1 - Although the concentration at the NOAEL was not measured, the lead investigator on this study said that he believes it was probably 0.4 mg/kg. 

BW = Body Weight 

C = Condition 

D = Deformities 

E = Enzymes 

FG = Feed:Gain ratio 

L = Length 

R= Reproduction 

S = Survival 

SA = Swimming Activity 
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In addition to the studies summarized in Table 4-2, the NRC (1980, 1994, 2005) provide 

summaries of additional dietary selenium toxicity studies on domestic chickens that were 

not available to the authors of this report. These data are presented in Tables 4-3 through 4-

5, respectively. 

Table 4-3 

Summary of additional selenium toxicity studies from the NRC (1980). 

Concentration (mg/kg) 

Age Chemical Form NOAEL LOAEL Effect Reference 

Adult Se in corn, barely, wheat 2.5 5 increased chick mortality Moxon (1937) 

Adult Na2SeO3 - 6.5 deformed embryos Moxon (1937) 

1 day Na2SeO3 4 8 decreased weight gain Moxon (1937) 

1 day selenous acid 2 8 increased chick mortality Thapar et al (1969) 

1 day SeO2 - 5 increased mortality Hill (1979) 

Adult Na2SeO3 - 8 embryo necrosis Gruenwald (1958) 

Table 4-4 

Summary of additional selenium toxicity studies from the NRC (1994). 

Age Chemical Form 

Concentration 

(mg/kg) Effect Reference 

Immature Na2SeO3 + Se in wheat 10 Reduced growth Carlson and Leitis (1957) 

Laying hen Se in wheat 10 Reduced hatchability Moxon and Wilson (1944) 

Table 4-5 

Summary of additional selenium toxicity studies from the NRC (2005). 

Concentration (mg/kg) 

Age Chemical Form NOAEL LOAEL Effect Reference 

Adult selenite 5 10 reduced growth Jensen and Chang (1976) 

1 day selenite 1 5 reduced growth Jensen (1986) 

For all of the studies on chickens presented above, it is unclear whether the dietary 

concentrations are provided as ww or dw. It is likely that the dietary concentrations are ww 

but that the %moisture is relatively low (e.g., 10%) as chicken feed is usually in the form of 

dry pellets. So, there is some uncertainty about the exact concentration of selenium in the 

studies above on chickens, but it is likely to be a small effect; e.g., 5 mg/kg-ww with a 

%moisture of 10% is equivalent to 5.6 mg/kg-dw. 

Eiders 

A single suitable study was identified (Franson et al. 2007). That study provided both a 

NOEAL and LOAEL for weight loss (i.e., a major effect) in diet, but did not examine 

reproductive success (Table 4-2). 

Kestrels 

Three studies have been conducted on the toxicity of dietary selenium to kestrels (Santolo et 

al. 1999, Yamamoto et al. 1998, Yamamoto and Santolo 2000). The first two studies (i.e., 

Santolo et al. 1999, Yamamoto et al. 1998) did not show an effect at up to 12 mg/kg-dw. In 

the third study (Yamamoto and Santolo 2000), the authors state that kestrels fed 12 mg/kgdw selenium had a lower “normalized body fat” than the kestrels in the other treatments. 
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However, the difference was not significant (i.e., p = 0.173) and was, therefore, excluded 

from Table 4-2 and consideration in the development of a screening value. 

Mallards 

Mallards are the bird species for which the most information is available. Fourteen suitable 

studies were identified on mallards. The studies examined effects on adults, ducklings, and 

reproductive success (Table 4-2). 

Pheasants 

A single suitable study was identified (Latshaw et al. 2004). That study provided a NOAEL 

and LOAEL based on mortality and reduced egg laying (Table 4-2).

Screech owls 

A single suitable study was identified (Wiemeyer and Hoffman 1996). That study provided 

both a NOAEL and LOAEL for selenium in the diet and eggs based on weight loss and 

reduced reproductive success. A NOAEL and LOAEL was also provided for minor effects; 

i.e., changes in enzyme concentrations (Table 4-2). 

4.3 IMPROVED PERFORMANCE STUDIES

Among birds that are raised on poultry farms, there has been extensive research in an 

attempt to balance the beneficial effects of selenium supplementation against the toxicity of 

selenium. These studies generally provide NOAELs only and do not necessarily examine the 

effects of selenium on reproductive success. The available studies are summarized in Table 

4-6 below. 

Table 4-6 

Studies evaluating improved performance in poultry from selenium supplementation. 

Effects 

Bird Study 

Life 

stage 

Dietary 

form 

NOAEL

(mg/kg) dw/ww Survival Weight 

Egg 

output 

Reproductive

success 

Chicken 24 Adult selenite 2.6 ? X 

Chicken 24 Adult Seyeast 2.9 ? X 

Chicken 27 1 day selenite 8.17 dw X 

Japanese 

quail 

3 1 day selenite 1.2 ? X X 

Japanese 

quail 

28 10 

day 

selenite 0.2 dw X 

The most sensitive endpoint for selenium effects in birds is reproductive success (i.e., egg 

hatchability and chick survival). However, since none of the performance studies examined 

the potential effects of selenium on reproductive success, these studies should not be used to 

put a lower bound on the acceptable level of dietary selenium for birds. 

4.4 DISCUSSION

More than 40 selenium toxicity studies were reviewed in an effort to determine dietary 

selenium screening values for species representative of birds in the North Bay (Table 4-1). 

Based on this review, species which are appropriate for use in developing screening criteria 

for the North Bay, and for which usable selenium toxicity data are available, include: 
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chickens (Gallus gallus), common eiders (Somateria mollissima), mallards (Anas 

platyrhynchos), pheasants (Phasianus colchicus), and screech owls (Otus asio). 

The data presented above (Tables 4-2 through 4-5; regardless of selenium type) were 

grouped into effects on immature birds (i.e., chicks and ducklings), adults, and reproductive 

success (i.e., egg hatchability, chick survival, egg fertility, etc.) and are presented in Table 4-

7 below. 

Table 4-7 

Summary of NOAELs and LOAELs (mg/kg-dw) reported in this section grouped by whether the 

effects were measured in chicks/ducklings, adult birds, or in terms of reproductive success. 

Studies using all selenium species are shown here. 

Chick/Duckling Adult Reproductive success 

NOAEL LOAEL 

Effect 

Threshold NOAEL LOAEL 

Effect 

Threshold NOAEL LOAEL 

Effect 

Threshold 

0.15 10 1.2 5 10 7.1 2.5 5 3.5 

0.2 15 1.7 3 5 3.9 

0.9 4.3 2.0 5 10 7.1 

1 5 2.2 - 6.5 - 

2 8 4.0 - 8 - 

4 8 5.7 - 8 - 

4.3 13.5 7.6 - 10 - 

- 5 - 

- 10 - 

Chicken 

SMCV 2.9 SMCV 7.1 SMCV 4.6 

Eider 20.6 57.7 34.5 

0.2 16.5 1.9 11.2 22.3 15.8 0.2 10.9 1.5 

0.2 16.9 1.9 13.8 33.8 21.6 0.37 6.5 1.6 

0.2 16.9 1.9 22.3 44.7 31.6 0.4 9.8 2.1 

10.9 21.6 15.3 0.6 10.6 2.6 

10.9 21.6 15.3 3.9 7.8 5.5 

16.5 65.4 32.9 4.6 9.0 6.4 

16.5 65.4 32.9 10.9 27.3 17.3 

16.9 66.9 33.6 

17.0 33.7 24.0 

Mallard 

SMCV 10.4 SMCV 22.1 SMCV 3.6 

Pheasant 0.4 9.3 1.9 

Screech 

owl 

8.8 30 16.2 

From the data presented in Table 4-7, it would appear that immature chickens are more 

sensitive to selenium than immature mallards, although there was some variation among 

chicken strains (see Table 4-2). The available studies on reproductive success appear to 

indicate that mallards are a little more sensitive than chickens but that screech owls are less 

sensitive than either chickens or mallards. For pheasants, the small amount of available data 

indicates that reproductive success is affected at relatively low concentrations. 

State of Utah 

The State of Utah is also deriving dietary selenium screening values protective of birds 

feeding at the Great Salt Lake (Ohlendorf 2007, provided as Appendix B). To develop a 

selenium screening value for the Great Salt Lake, the Great Salt Lake Science Panel 

compiled the concentration of selenomethionine in feed, total selenium in eggs, and mean 

hatching success per treatment from 6 studies on mallards (i.e., Heinz et al. 1987, 1989, 

Heinz and Hoffman 1996, 1998, Stanley et al. 1994, 1996; all of these studies are included 

in Tables 4-1, 4-2, and 4-7) and examined the data using both a logistic regression and a 
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hockey-stick regression. The regressions were used to derive concentrations at which 10% 

of the population would be affected (EC10) for both dietary and egg selenium 

concentrations. The results for mean hatching success are as follows (mg/kg-dw): 

! Dietary selenium 

o Logistic regression: 4.9 

o Hockey stick-regression: 4.4 

! Egg selenium 

o Logistic regression: 12.5 

o Hockey stick-regression: 11.5 

Figure 4-1 Logistic regression of mallard dietary selenium concentration vs. proportion of 

eggs hatching. From Ohlendorf (2007). 
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Figure 4-2 Hockey stick regression of mallard dietary selenium concentration vs. proportion 

of eggs hatching. From Ohlendorf (2007). 

Figure 4-3 Logistic regression of mallard egg selenium concentration vs. proportion of eggs 

hatching. From Ohlendorf (2007). 
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Figure 4-4 Hockey stick regression of mallard egg selenium concentration vs. proportion of 

eggs hatching. From Ohlendorf (2007). 

The logistic regressions were previously reported in Ohlendorf (2003) and the hockey-stick 

regression for egg selenium was previously reported in Adams et al. (2003). The paper by 

Adams et al. (2003) is the third paper published by the same group of authors on this topic; 

i.e., Fairbrother et al. (1999, 2000) used the data from fewer studies (i.e., Heinz et al. 1989 

and Stanley et al. 1994) to derive an EC10 for eggs of 16 mg/kg-dw. 

As discussed above for the regression approach used by Beckon (2007) to analyze fish data, 

the State of Utah (Ohlendorf 2007) performed regressions on the treatment means, instead of 

the raw data from each individual, which artificially reduces the variation; i.e., the 

confidence limits on the regression lines in Figures 4-1 through 4-4 may be much larger. 

Further, the shape of the relationship may change when using the raw data. Thus, the EC10 

and EC20 presented in Figures 4-1 through 4-4 should be viewed as approximations. Lastly, 

the data used in the logistic regressions (Ohlendorf 2003) was also normalized against the 

controls. This may have obscured a biphasic relationship (see Beckon et al. (2008)). 

There are a few minor problems with the data used by the State of Utah (Ohlendorf 2007) 

that should be kept in mind. First, the concentration of selenomethionine used in the logistic 

regressions was assumed by the State of Utah (see Ohlendorf 2003) to be dry weight 

(Ohlendorf 2007) but the concentrations were actually presented as wet weight. The State of 

Utah (Ohlendorf 2007) acknowledged this by stating that the values derived in the logistic 

regressions for dietary concentrations should be increased by 11 percent (i.e., to 5.4 mg/kgdw). Further, the State of Utah (see Ohlendorf 2003) used the “nominal” selenium 

concentrations added to feed in their analyses and not the measured concentrations. This 

adds an additional layer of uncertainty to the thresholds reported by the State of Utah 

(Ohlendorf 2007) as the actual concentration can differ from the nominal concentration 

either positively or negatively, depending upon the individual treatment. 
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Other selenium screening values 

The National Irrigation Water Quality Control Program (NIWQP) recommends a lower 

screening value for selenium in bird eggs than the State of Utah; i.e., 6 mg/kg-dw (NIWQP 

1998). The USGS (Seiler et al. 2003) used logistic regressions to derive threshold 

concentrations for selenium induced teratogenicity (a less sensitive endpoint than hatching 

success) in bird eggs collected from several selenium-impacted field sites across California 

(Table 4-8). The NIWQP also recommends a screening value for dietary selenium of 3 

mg/kg-dw. Presser and Luoma (2006) also reviewed the toxicity literature and concluded 

that dietary “concern levels range from 2 to 7” mg/kg-dw for birds. 

Table 4-8 

Selenium screening values for teratogenesis in bird eggs (mg/kg-dw) (Seiler et al. 2003). 

Bird EC01 EC10 

Ducks 15 23 

Black-necked stilts 14 37 

American avocets 41 74 

Allometric adjustment for dietary concentrations 

The species of concern in the North Bay identified by USFWS (2007) have not been 

examined in selenium toxicity studies. One approach for applying screening values 

developed for one species to another species is to allometrically adjust the screening values 

(Sample and Arenal 1999). This adjusts for differences in body weight, metabolism, 

pharmacokinetics, and sensitivity to provide the best available estimates of species-specific 

toxicity. The equations use to perform the allometric adjustments, and parameter values, are 

provided in Appendix C. 

Allometric adjustments were applied to predict dietary selenium screening values for the 

species of concern in the North Bay identified by USFWS (2007) from three example 

dietary toxicity values protective of reproductive success: 

1. the EC10 for mallards fed selenomethionine derived using a hockey-stick 

regression by Adams et al. (2003) and used by the Great Salt Lake Science 

Committee (Ohlendorf 2007); 

2. the effect threshold (Table 4-2) derived from the study by Ort and Latshaw (1978) 

for chickens, and; 

3. the effect threshold (Table 4-2) derived from the study by Wiedemeyer and 

Hoffman (1996) for screech owls. 

These studies were chosen as the most representative of the available toxicity data although 

other toxicity values may also be used (e.g., the SMCV for chicken reproductive studies). 

However, since reproductive success is the most sensitive endpoint in birds, it is 

recommended that the allometrically adjust dietary threshold all be based on that endpoint. 

The allometrically adjusted screening values are presented below in Table 4-9. 
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Table 4-9 

Allometrically adjusted dietary selenium screening values. 

Dietary Screening Value (mg/kg-dw) Allometrically Adjusted from 

Bird species Mallarda

 Chickenb Screech Owlc

Bald eagle 9.7 4.0 104.3 

California clapper rail 2.2 0.9 - 

Greater scaup 3.9 1.6 - 

Lesser scaup 3.2 1.3 - 

White-winged scoter 5.6 2.3 - 

Surf scoter 4.1 1.7 - 

Black scoter 3.9 1.6 - 

Notes: 

a - EC10 for reduced hatching success from Adams (2003) and Ohlendorf (2007) of 4.4 mg/kg-dw 

b - effect threshold for reduced hatching success of 3.9 mg/kg-dw from Ort and Latshaw (1978) (Table 4-2) 

c - effect threshold for reduced hatching success of 17.3 mg/kg-dw from Wiemeyer and Hoffman (1996) (Table 4-2) 

Note that allometric adjustments are used to adjust dietary screening values (Sample and 

Arenal 1999) and are not used to adjust screening values for selenium in eggs. Further, there 

is some uncertainty associated with the use of allometric scaling as the allometric scaling 

factor used here was developed for acute toxicity data its “...applicability to chronic toxicity 

data is unknown. The modes of action for acute and chronic effects differ for many 

chemicals. As a consequence, it is likely that scaling factors based on chronic toxicity data 

will also differ from those based on acute toxicity data” (Sample and Arenal 1999). It should 

also be noted that the allometric scaling factor used here is not specific to selenium, but a 

mean value for other contaminants. 

Allowable limits in bird feed 

The FDA has approved the use of sodium selenate, sodium selenite, and selenium yeast at 

up to 0.3 mg/kg for chickens, swine, turkeys, sheep, cattle, and ducks (21CFR573.920). The 

major (i.e., 60-80%) form of selenium in selenium yeast is selenomethionine (Capelo et al. 

2004, McSheehy et al. 2005, Polatajko et al. 2005). 

Nutritional requirements 

Selenium is a micro-nutrient. When animals do not have enough selenium in their diet, 

selenium deficiency results; i.e., poor growth, muscular dystrophy, and chick mortality 

(Ullrey 1992, NRC 1994). Due to the economic importance of the poultry industry, the 

nutritional requirements to maintain an adequate growth rate and normal metabolism of 

some of the more widely raised species have been established by the NRC (1994; see Table 

4-10 below). 

Table 4-10 

Selenium nutritional requirements for commercially reared birds (mg/kg-dw). 

Age (weeks) 

Species Strain 0-2 2-4 4-6 6- 

Chicken Leghorn white egg-laying 0.17 0.17 0.17 0.17 

Chicken Leghorn brown egg-laying 0.16 0.16 0.16 0.11 

Chicken Broiler 0.17 0.17 0.17 0.17 

Turkey - 0.22 0.22 0.22 0.22 

Duck - 0.22 - - - 

Quail Japanese 0.22 0.22 0.22 0.22 

Source: NRC (1994) 
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Maximum tolerable limits 

The NRC (2005) defines the maximum tolerable level as the concentration “...of a mineral 

is the dietary level that, when fed for a defined period of time, will not impair animal health 

and performance.” The NRC (2005) has set the following maximum tolerable levels for 

selenium in bird diets: a) 3 mg/kg for poultry and b) 2 mg/kg-dw for aquatic birds. Poultry 

includes chickens, turkeys, grouse, and pheasants, as well as ducks and geese (which are 

water birds). It should also be noted that the studies on kestrels did not produced adverse 

effects at up to 12 mg/kg-dw (Santolo et al. 1999, Yamamoto et al. 1998, Yamamoto and 

Santolo 2000). 

Additional considerations 

Issues to be considered in the collection and interpretation of samples include: 

! tissue concentrations in fishes may vary with the time of year sampled 

! tissue concentrations may vary by geographic location 

! if sampling dietary items of birds, selenium concentrations may also differ among 

their prey items 

! number of samples to collect 

! the frequency of collections 

! whether samples should be composited 

! how to compare measured concentrations to screening values 

These issues are discussed in greater detail in Section 6. 

A point worthy of consideration is whether the type of food consumed affects the 

bioavailability and/or toxicity of selenium. The bird toxicity studies reviewed above would 

likely have used a feed that mainly consisted of cereal grains and oilseed meals (NRC 1994), 

whereas many of the bird species of concern in the North Bay are at least partially 

carnivorous. Among pigs, “animal Se sources generally have a lower (<25%) absorption 

rate than Se from plant tissue. However, Se absorption is influenced by the type of animal 

product being fed. For example, the absorption of Se from bovine milk is high and 

equivalent to that of selenite, whereas Se in meat and bone meal or poultry by-products is 

"20%” (Mahan 2001). This implies that the bioavailability of selenium is likely to be lower 

in the diets of the birds in the North Bay than in the diets used in the toxicity studies 

reviewed here. Thus, dietary screening levels for selenium may need to be adjusted upwards 

when they are based on studies where the test animal was fed mainly vegetable matter. 

The toxicity values presented above were not derived using the populations and species of 

birds that occur in the North Bay nor using the environmental conditions in the North Bay. 

This means that there is some uncertainty as to whether any screening criteria are actually 

applicable. Similarly, McDonald and Chapman (2007) recommend that screening values 

only be used to determine if further evaluation is necessary. They recommend that if sample 
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concentrations exceed an appropriate screening value that actual impacts on the species of 

concern be assessed. Since reproductive success is the most sensitive endpoint in birds, it 

may be relatively easy to determine whether reproductive success of aquatic birds is being 

adversely impacted in the North Bay. This would require that an egg monitoring program be 

designed and implemented. 

Future research and data gaps 

Several data gaps and areas where further research may be beneficial to the Water Board 

were identified during the evaluation of selenium toxicity to birds in the North Bay, as 

follows: 

1. The logistic regressions developed by the State of Utah (Ohlendorf 2007) to derive 

a dietary effect threshold for selenium should be redone using the correct data. 

2. Toxicity studies have not been performed on most of the species of concern in the 

North Bay (USFWS 2007). Thus, this report has assumed that toxicity studies on 

mallards, chickens, pheasants, and screech owls are representative of the species of 

concern. The species of concern may be more or less sensitive. 

3. If toxicity testing is not performed for the species of concern in the North Bay, 

another approach may be to develop selenium-specific allometric scaling factors for 

organic selenium in avian diets. 

4. Species-specific regressions to convert dietary selenium concentrations to egg 

concentrations are lacking. If the Water Board decides to use the protection of egg 

hatching as a numeric criterion, this will be an important data gap. 

5. The selenium toxicity studies performed to date have used selenite, 

selenomethionine, and selenized yeast. The degree to which the dietary 

formulations used in the toxicity studies reflects the speciation of selenium 

speciation in the food of birds of concern in the North Bay is not very well known. 

Since the speciation of selenium affects its toxicity, this may be an important 

consideration. 

6. Toxicity experiments are uneven in assaying the actual amount of selenium in the 

experimental feeds. Many studies only report the estimated concentration added. 

Thus, the reported amount of selenium provided to birds in selenium toxicity 

studies is subject to some error. The error can be either positive or negative, 

depending upon the individual treatment. 

7. Many toxicity studies performed to date report their concentration data in wet 

weight and do not provide a %moisture. Although a default %moisture can be used 

to estimate dry weight concentrations, this adds an additional layer of uncertainty 

as the actual %moisture will vary from sample to sample, at the very least. 

8. A detailed re-analysis of the published toxicity data using, for example, biphasic 

regressions (Beckon 2007, Beckon et al. 2008) would provide a greater level of 

confidence in estimates of the threshold for toxic effects of selenium vs. a LOAEL, 

NOAEL, or effect threshold. However, if only the treatment means are available for 

the analyses, it should be remembered that this type of analysis will mask the true 

variability in the data. 
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4.5 CONCLUSIONS

Exposure to elevated levels of selenium can cause adverse effects in birds. Raptors, such as 

screech owls and kestrels, appear to be less sensitive to dietary exposures to selenium than 

ducks and chickens. While it has generally been recognized that reproductive success is the 

most sensitive endpoint in mallards, a review of the toxicity data available for chickens 

indicates that the growth/survival of young is a more sensitive endpoint in chickens. Thus, it 

is an open question as to which endpoint will be more sensitive for the birds of concern in 

the North Bay. 

Numerous selenium screening values protective of birds have been proposed. For dietary 

selenium, the recommended screening values are usually in the range of 3-5 mg/kg-dw. For 

bird eggs, the range of recommended screening values is considerably larger, ranging from 6 

to 74 mg/kg-dw. 

Based upon the data presented here, there are several options for selenium screening values 

protective of birds. Those options are as presented below (all in units of mg/kg-dw). All of 

the screening values are well above the dietary requirements for birds of approximately 0.2 

mg/kg-dw (Table 4-10), with the exception of some of the allometrically adjusted dietary 

thresholds, which are close to the dietary requirements. 

1. The lowest dietary SMCV of 2.9 (Table 4-7) 

2. Lowest dietary concentration at which adverse effects were reported (Table 4-7) 

a. LOAEL of 4.3 

3. The effect concentrations protective of reproductive success in mallards derived by the 

State of Utah (Ohlendorf 2007): 

a. Dietary selenium 

i. Logistic regression: 5.4 

ii. Hockey stick-regression: 4.4 

b. Egg selenium 

i. Logistic regression: 12.5 

ii. Hockey stick-regression: 11.5 

4. Allometrically adjusted dietary thresholds for the species of concern in the North Bay 

(Table 4-9) 

a. Note that allometric adjustment predicts that smaller birds will have lower dietary 

thresholds, with the lowest being approximately 1 mg/kg-dw for California clapper rails 

The choice of using an SMCV, LOAEL, effect threshold, an EC01, or an EC10 is dependent 

upon both the level of risk and uncertainty the Water Board wishes to accept. This issue was 

presented in more detail in the Introduction. It is worth noting that of these values, only the 

LOAEL represents concentrations at which effects have been demonstrated. However, since 

only a few concentrations have been tested, effects may actually occur at lower 

concentrations.
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5 MARINE MAMMAL TOXICITY REVIEW

Although some whale and dolphin species may occasionally enter the waters of the North 

Bay, they are generally considered infrequent visitors. The three marine mammal species 

that are considered to be resident in the North Bay include (DFG 2007, USFWS 1992): 

! River otter (Lontra canadensis) 

! Harbor seal (Phoca vitulina) 

! California sea lion (Zalophus californianus) 

Radio telemetry data indicates that harbor seals and sea lions venture in and out of the North 

Bay (USFWS 1992). None of the three marine mammal species found in the North Bay are 

special status species. 

Among mammals exposure to elevated levels of selenium can cause reduced growth, 

emaciation, hair loss, lesions, paralysis, reduced conception rates, reduced litter size, 

increased stillbirths, and mortality (Mahan 2001, NRC 1980, 2005, Ohlendorf 2003). When 

selenium concentrations are high enough in a large enough area, these effects can lead to 

population declines for the most affected species. 

The step in the effort to determine whether marine mammals may be adversely impacted by 

selenium in the North Bay is to evaluate the toxicity of selenium to the marine mammals. 

Next, potential selenium screening values for the species in the North Bay are proposed. 

Lastly, the screening values are compared to nutritional requirements. 

5.1 METHODS

No toxicity studies have been performed on the effects of selenium on any marine mammal. 

Among mammals, selenium toxicity studies have been performed on cows, dogs, goats, 

hamsters, mice, pigs, and rats. However, “due to the reduction of selenite and selenate, and 

the formation of insoluble particles in the rumen, cattle and sheep have [a] lower absorption 

of selenium... than non-ruminant species” (NRC 2005). Therefore, toxicity studies where 

selenium has been fed to cows, sheep, and goats are probably not representative of marine 

mammals. Herbivorous and semi-herbivorous rodents (e.g., hamsters, mice, and rats) have 

reducing conditions in their intestines (as opposed to the stomach of ruminants) and also 

engage in coprophagy. Therefore, dietary toxicity studies on hamsters, mice, and rats are not 

likely to be reflective of carnivores, including marine mammals. Thus, it was assumed that 

dietary toxicity studies on pigs and dogs are likely to be representative of marine mammals. 

Dietary selenium toxicity studies using pigs have been reviewed by the NRC (1980, 2005) 

and Mahan (2001). Studies on dogs have also been reviewed by the NRC (1980). The 

studies included in those reviews are summarized below in the Results section. 

5.2 RESULTS

As stated above, no selenium toxicity studies on marine mammals were found. However, it 

was assumed that dietary toxicity studies on pigs and dogs are representative of marine 
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mammals as all three groups are carnivorous mammals. The studies on pigs and dogs 

reviewed by the NRC (1980, 2005) are provided in Tables 5-1 and 5-2 below, respectively. 

Table 5-1 

Dietary toxicity studies on pigs from NRC (1980, 2005). 

Concentration (mg/kg) 

Age Chemical Form NOAEL LOAEL Effect Reference 

"growing" Na2SeO3 4 8 reduced weight gain Goehring et al. (1984) 

"growing" Na2SeO3 5 15 reduced weight gain Mahan and Magee 

(1991) 

8-10 wk Na2SeO3 and SeMet 0.4 25 reduced weight gain, 

paralysis 

Panter et al. (1996) 

24.7 kg Na2SeO3 or Se-yeast 5 10 reduced weight gain Kim and Mahan 

(2001a) 

Adult Na2SeO3 4 8 offspring with reduced 

weight 

Poulsen et al. (1989) 

25 kg Na2SeO3 or Se-yeast 3 7 lower number of live 

born 

Kim and Mahan 

(2001b) 

5 kg Na2SeO3 and SeMet 10 20 reduced weight gain Herigstad et al. (1973) 

35 kg Seleniferous corn 5 10 toxicosis Schoening (1936) 

15 kg Na2SeO3 - 7 reduced weight gain Wahlstrom et al. (1956) 

14 kg Na2SeO3 - 10 reduced weight gain Wahlstrom et al. (1956) 

13 kg Na2SeO3 - 11 reduced weight gain Wahlstrom et al. (1956) 

15 kg Na2SeO3 - 10 reduced reproductive 

success 

Wahlstrom and Olson 

(1959) 

16-19 kg Na2SeO3 - 24 Anorexia, death Miller and Schoening 

(1938) 

Table 5-2 

Dietary toxicity studies on dogs from NRC (1980). 

Concentration (mg/kg) 

Age Chemical Form NOAEL LOAEL Effect Reference 

60 days Seleniferous corn - 7.2 Decreased weight gain Rhian and Moxon (1943) 

150 days Na2SeO3 - 10 Decreased weight gain Rhian and Moxon (1943) 

72 days Seleniferous corn - 20 Decreased weight gain Rhian and Moxon (1943) 

"young" Na2SeO3 - 20 Decreased weight gain Moxon (1937) 

"young" Seleniferous corn - 20 Decreased weight gain Moxon (1937) 

For all of the studies on presented above, it is unclear whether the dietary concentrations are 

provided as ww or dw. Thus, the units are given only as mg/kg. 

Mahan (2001) summarizes the existing studies on pigs by stating that “chronic selenosis 

generally occurs when diets or feedstuffs contain 5 to 20 ppm Se.” 

5.3 DISCUSSION

No selenium toxicity studies have been performed on marine mammals. Therefore, it was 

assumed that toxicity studies on other mammals would be representative of marine 

mammals. Among the mammals that have been studied, it was assumed that pigs and dogs 

are more likely to be representative of the marine mammals of concern in the North Bay 

than studies using ruminants and rodents. This is because ruminants and rodents have 

reducing conditions in their gastro-intestinal tracts that may reduce the bioavailability of 

selenium relative to the gastrointestinal tracts of pigs, dogs, seals, sea lions, and otters. 
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The studies reviewed by the NRC (1980, 2005) indicate that effects in pigs were not 

observed at dietary concentrations of less than 7 mg/kg (Table 5-1). In Nutrient 

Requirements of Swine (NRC 1998), a study that was not reviewed in the Mineral Tolerance 

of Animals (NRC 1980, 2005) was cited as causing adverse effects in pigs at 5 mg/kg. 

Mahan (2001) also states that effects have been observed in pigs at feed concentrations at 5 

mg/kg. Thus, it appears that the lowest reported dietary concentration at which selenium has 

been shown to affect pigs is approximately 5 mg/kg. 

The studies reviewed by the NRC (1980) on dogs did not administer dietary concentrations 

below 7.2 mg/kg. Since adverse effects were observed at that dietary concentration, it is not 

possible to determine the an actual effect threshold in dogs. 

A point worthy of consideration is whether the type of food consumed affects the 

bioavailability and/or toxicity of selenium. The pig studies reviewed above would likely 

have used a feed that mainly consisted of corn and soybeans (NRC 1998), whereas seals, sea 

lions, and river otters are entirely carnivorous. Among pigs, “animal Se sources generally 

have a lower (<25%) absorption rate than Se from plant tissue. However, Se absorption is 

influenced by the type of animal product being fed. For example, the absorption of Se from 

bovine milk is high and equivalent to that of selenite, whereas Se in meat and bone meal or 

poultry by-products is "20%” (Mahan 2001). This implies that the bioavailability of 

selenium is likely to be lower in the diets of marine mammals than pigs and that the dietary 

screening level for selenium may need to be adjusted upwards of 5 mg/kg for marine 

mammals. 

Allowable limits in feed 

“In 1982, the FDA approved the addition of 0.3 ppm of selenium to diets for pigs up to 20 

kg, because 0.1 ppm of added selenium does not always prevent deficiency signs in weanling 

pigs. The current regulation allows up to 0.3 ppm of selenium in the diet for all pigs” (NRC 

1998). 

Nutritional requirements 

It should also be remembered that selenium is a micro-nutrient. When mammals do not have 

enough selenium in their diet, selenium deficiency results; i.e., edema, hepatic necrosis, poor 

growth, white muscle disease, mulberry heart disease, impaired reproduction, reduced milk 

production, impaired immune response, and mortality (NRC 1998). Due to the economic 

importance of the pig farming industry, the nutritional requirements to maintain an adequate 

growth rate and normal metabolism of pigs have been established by the NRC (1998; see 

Table 5-3 below) and Mahan (2001). 

Table 5-3 

Selenium nutritional requirements for commercially reared pigs (mg/kg-dw) by body weight (kg). 

NRC (1998) Mahan (2001) 

Body weight (kg) 3-10 10-20 20-120 125+ 5-20 20-60 60+ - 

growing pigs 0.33 0.28 0.17 0.30 0.20 0.15 

gestating female 0.17 0.3 

lactating female 0.17 0.3 

Sexually active boars 0.17 
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Maximum tolerable limits 

The NRC (2005) defines the maximum tolerable level as the concentration “...of a mineral 

is the dietary level that, when fed for a defined period of time, will not impair animal health 

and performance.” The NRC (2005) has set a maximum tolerable level of selenium in feed 

for pigs of 4 mg/kg (wet weight vs. dry weight not specified). The NRC (1998) also 

provides the following discussion of the maximum tolerable limit for selenium in pig feed: 

“When fed to growing swine as sodium selenite, sodium selenate, selenomethionine, or 

seleniferous corn, selenium does not produce toxicity at levels of less than 5 ppm. In some 

cases, however, a level of 5 ppm and levels from 7.5 to 10 ppm have produced toxicity. Signs 

of toxicity include anorexia, hair loss, fatty infiltration of the liver, degenerative changes in 

the liver and kidney, edema, occasional separation of hoof and skin at the coronary band, 

and symmetrical, focal areas of vacuolation and neuronal necrosis.” 

Future research and data gaps 

Several data gaps and areas where further research may be beneficial to the Water Board 

were identified during the evaluation of selenium toxicity to marine mammals in the North 

Bay, as follows: 

1. Toxicity studies have not been performed on any of the marine mammals that 

reside in the North Bay. Thus, this report has assumed that toxicity studies on pigs 

and dogs are representative of harbor seals, sea lions, and river otters. The marine 

mammals in the North Bay may be more or less sensitive than pigs and dogs. 

2. Toxicity studies on pigs have used a diet based on vegetable matter. The 

bioavailability of selenium in such diets is higher than in animal matter based diets. 

This means that the bioavailability of selenium in the diets of marine mammals 

may be lower than the in the toxicity studies that have been performed on pigs. 

3. For this report, only the secondary literature was reviewed. A thorough 

examination of the primary literature (as was done for fish and birds) may prove 

more informative. 

4. A detailed re-analysis of the published toxicity data for pigs using, for example, 

biphasic regressions (Beckon 2007, Beckon et al. 2008) would provide a greater 

level of confidence in estimates of the threshold for toxic effects of selenium vs. a 

LOAEL, NOAEL, or effect threshold. However, if only the treatment means are 

available for the analyses, it should be remembered that this type of analysis will 

mask the true variability in the data. 

5. For harbor seals and sea lions, the collection and analysis of blood, hair, and 

blubber samples appears to be fairly routine. These types of samples can be 

collected without killing the seal/sea lion. It might prove useful to examine the 

dietary toxicity literature for pigs and develop relationships between dietary 

selenium levels and blood, fat, and/or hair levels in pigs that may be used to 

evaluate data collected from seals and sea lions in the North Bay. 

6. The selenium toxicity studies performed to date have used selenite, 

selenomethionine, and selenized yeast. The degree to which the dietary 

formulations used in the toxicity studies reflects the speciation of selenium 
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speciation in the food of marine mammals of concern in the North Bay is not very 

well known. Since the speciation of selenium affects its toxicity, this may be an 

important consideration. 

7. The secondary reviews do not detail whether the toxicity studies performed on pigs 

measured the selenium concentration in feed as ww or dw. This adds an additional 

layer of uncertainty. 

5.4 CONCLUSIONS

Exposure to elevated levels of selenium can cause adverse effects in mammals. Based upon 

the data presented here, there are several options for selenium screening values protective of 

marine mammals. Those options are as follows (all in units of mg/kg): 

1. The maximum tolerable limit recommended by that NRC (2005): 4 

2. Lowest concentration at which adverse effects have been reported: 5 

These proposed screening values are well in excess of the dietary requirements for pigs 

(which are assumed here to be representative of marine mammals) of approximately 0.3 

mg/kg-dw (Table 5-3). 

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6 ISSUES AFFECTING THE CHOICE OF SCREENING VALUES

This report has presented several options for screening values protective of different 

receptor groups. Here, we discuss issues related to selecting a screening value for use in the 

TMDL process. 

Derivation of screening values 

In Section 1, multiple methods were identified that have been used in the toxicity literature 

to derive screening values, including: 

! LOAELs 

! Effect thresholds 

! Species mean chronic values 

! EC01 or EC10 

! Species sensitivity distributions 

Further, there are multiple means of calculating an effect concentrations (i.e., an EC01 or 

EC10) or a species sensitivity distribution. This means that different mathematical 

techniques can be applied to the same data to derive different screening values. 

Determining which method is the best is no easy matter, as each method has its own 

limitations. Most of the selenium toxicity data examined in this review was collected from 

experiments designed to be analyzed using an ANOVA; i.e., there are several groups of 

animals in an experiment and each group is exposed to a different selenium dose. These 

types of experiments produce a NOAEL and a LOAEL; i.e., the highest selenium dose at 

which no adverse effects were observed and the lowest selenium dose at which adverse 

effects were observed. 

Some authors have re-analyzed the treatment means from these kind of experiments using a 

regression analysis (e.g., logistic, hockey stick, or biphasic) to obtain a better understanding 

of the threshold for toxic effects than is provided by the NOAEL and LOAEL (e.g., Adams 

2003, Ohlendorf 2003, 2007, Beckon 2007, 2008). However, in analyzing only the treatment 

means, significant variation is lost and the toxicity thresholds derived are uncertain. 

Toxicity experiments designed to be analyzed using regression analyses have the potential to 

be much more powerful than toxicity tests designed to be analyzed using an ANOVA; i.e., 

“the purposes of accurately determining risk and maximizing benefit are more efficiently 

served by the allocation of limited experimental resources to a greater range of dose levels 

rather than to multiple replicates at a small number of dose levels. The latter, a legacy of 

traditional hypothesis testing, is relatively poorly suited for fitting and comparing 

mathematical models of the dose-response relationship” (Beckon et al. 2008). Therefore, it 

is recommended that future toxicity studies be designed with a regression analysis in mind. 
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When using an effect concentration (EC) or a species sensitivity distribution, one must also 

decide on what level to protect in order to derive the screening value. For effect thresholds, 

it is common to set the threshold at a 1, 10, or even 20% effect. For species sensitivity 

distributions, it is common to choose a value that is protective of 95% of the species 

evaluated. Choosing what threshold to use should be related to what level may actually 

impact populations in the field, yet information on that is scant. Further, when comparing 

screening values derived using either of these methods, one must also consider the 

confidence limits; e.g., is an EC10 of 1.8 significantly different from an EC20 of 2.5? 

Uncertainties in toxicity testing 

There are some additional sources of uncertainty associated with toxicological studies 

reviewed here that should be considered when interpreting the results to set regulatory 

standards, including the following: 

! Responses to selenium differ among species and life-stage within the same 

species. However, results for one species of life-history stage are used to assess 

the effects on other species and life stages. For example, the only toxicity data 

available for white sturgeon were derived from short-term chronic (8-week) 

exposures to juvenile fish, but this fish is long-lived and does not reach sexual 

maturity until they are 10 years of age or more. 

! In short-term toxicity tests performed on fish in the laboratory, the concentrations 

in the tissue may not have reached steady state. Thus, screening values derived 

from short term studies where tissue concentrations have not yet reached 

equilibrium may underestimate the tissue level that causes toxicity. However, 

current recommendations are to use fish tissue concentrations as screening values 

(Hamilton 2002, 2003, 2004, USEPA 2004). 

! There is a lack of data linking the results of laboratory toxicity data to selenium 

concentrations and effects observed in the field. Even though effects have been 

demonstrated in the laboratory, it is unknown whether these effects will translate 

to population level effects in the field; i.e., population declines. 

Current selenium toxicity testing efforts 

At a recent meeting of the North American Industry Selenium Working Group (November 

16, 2007) several new selenium toxicity studies were announced. These studies include the 

following: 

! J.R. Simplot will spend approximately $500K for selenium toxicity testing with 

brown trout, Yellowstone cutthroat trout, and possibly a third species. 

! USEPA has performed a repeat of the bluegill toxicity test done by Lemly 

(1993a) upon which the USEPA (2004) based its draft ambient water quality 

criterion. Toxicity was found at higher concentrations that observed by Lemly 

(1993a). These results have not yet been published. 
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! The North American Industry Selenium Working Group may be funding 

relatively small studies on “site-specific assessments of selenium bioaccumulation 

in aquatic systems” and “tissue endpoint assessment.” However, the details of 

these studies are unknown. 

! Cameco is performing “water bird research in Saskatchewan including assessing 

food, productivity, and egg [Se] for tree swallows.” 

The research that is currently in progress described above is all on freshwater ecosystems. 

So, although the research is of general interest, it will not address the issues of greatest 

concern to the Board; i.e., the toxicity of selenium of brackish/marine ecosystems and the 

toxicity of selenium to the species of concern in the North Bay. Further, the current research 

does not appear to have been designed for use with regression analyses, meaning the 

NOAELs and LOAELs will be provided and deriving an EC20 will still be associated with 

some uncertainty. 

Conclusions 

Ultimately, given importance of selenium to the water resources of California, and the 

history of its effects, it is surprising that so little toxicity data on the species of concern in 

California is available. The methods used in selenium toxicity testing to date do not provide 

regulatory entities with a high degree of confidence in deriving toxicity thresholds and/or 

screening values. Lastly, field validation of selenium toxicity studies should also be 

performed to increase confidence in any screening levels derived from the existing toxicity 

literature. 

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7 ISSUES AFFECTING IMPLEMENTATION

When sampling invertebrates, fish, bird eggs, or other biological media to compare to the 

selenium screening values proposed in the earlier sections of this report, there are several 

issues that should be considered to maximize the use of the data. This section of the report 

provides a brief discussion of some of the more important issues. Some of these issues have 

been considered by various authors (e.g., Lemly and Skorupa 2007, McDonald and 

Chapman 2007, USEPA 2000, 2006). 

Spatial variability 

The concentration of selenium in the waters of the North Bay is expected to differ depending 

upon where samples are collected. Near selenium inputs (e.g., refineries, some creeks, and 

the San Joaquin River), selenium concentration in the water column are expected to be 

higher. This should result in higher concentrations for the biota that inhabit that area, 

providing that they have a relatively small range. For example, a clam may move only a few 

feet within its lifetime, but a sea lion may travel thousands of miles. Therefore, if there is an 

area where there is a localized input of selenium, elevated selenium levels would be 

expected in biota with a small range (e.g., the clam) but not necessarily in biota with a large 

range (e.g., the sea lion). This suggests that samples should be collected from areas near 

known selenium inputs from biota that are most likely to be impacted. To determine the 

extent of impacts, a number of locations should be sampled, not just areas near selenium 

inputs. 

Background 

Since selenium is a naturally occurring element, it is expected to be present at some 

concentration in all waters, soils, and sediments in the North Bay. Therefore, in addition to 

examining the spatial variability of selenium in the North Bay, it may be desirable to 

establish an area (or areas) that are representative of unimpacted (or reference) conditions. 

Then, biota may be sampled from that area and selenium concentrations in the tissues of 

biota from the background location can be compared to the samples collected from the 

North Bay to determine what areas are impacted. If this technique is used, it is advised that 

statistically robust techniques be used to compare sample locations, such as the t-test, 

ANOVA, or Kruskall-Wallis test. 

Temporal variability 

Concentrations of selenium in tissues may vary with the time of year. USEPA (2006) 

recommends that fish samples be collected during the same time of year during each sample 

event to avoid potential temporal effect. 

Comparing measured concentrations to screening values 

USEPA (2006) ”... recommends using the [one sample] t-test to determine whether the 

mean concentration of mercury in composite fish tissue samples exceeds the screening 

value. This involves a statistical comparison of the mean of all fish tissue data to the 

criterion. EPA recommends that this procedure also be used for determining impairment.” 

This technique is also applicable to selenium (without the compositing). To use a onesample t-test, the screening value is compared against the mean concentration for all of the 

biota samples of the same species collected from the location/time being evaluated. The 
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results of the test can be used to determine whether the measured concentrations 

significantly exceed the screening value or not. 

Another approach that can also be used is to compare the 95% upper confidence limit on the 

mean (UCL95) to the screening value. If the calculated UCL95 does not exceed the screening 

value, then no action is necessary. USEPA (2007b) provides free statistical software that can 

be used to calculate a UCL95s for selenium data in biological samples. 

Lastly, it is also possible to simply compare measured concentrations to the screening 

values. This method is likely to result in more exceedances than using a one sample t-test or 

a UCL95. In that respect, comparing individual concentrations to the screening value may be 

more “sensitive” but it is not statistically valid. 

Food webs 

It has been demonstrated that among the fish species in the North Bay that those that feed on 

clams, or are part of a food web that starts with clams, have higher selenium tissue 

concentrations (Stewart et al. 2004). This is because the concentrations of selenium in one of 

the more abundant clams in the North Bay, Potamocorbicula amurensis, are much higher 

than many of the other invertebrates that fish feed on (Stewart et al. 2004). Similarly, it 

would be expected that birds that feed on this species of clam have higher selenium 

exposures than other bird species. In addition, it is generally expected that fish and birds that 

feed on biota in the sediments of the Bay would have greater selenium exposures than fish 

and birds that feed in the open water column; i.e., the Delta smelt only feed on zooplankton 

near the top of the water column and would, therefore, be expected to be exposed to less 

selenium than fish that feed in the sediments, such as sturgeon. However, it is possible that 

species that have a lower potential for exposure may be more sensitive to selenium. Thus, 

the other species in the North Bay should not be ignored. 

Size of specimens 

Data collected by the USGS indicate that as the size of Sacramento splittail increases over 

their lifetime, the fish’s dietary habits change and that this affects their bioaccumulation of 

selenium. In addition, the concentrations of many metals increase with the size of the 

specimen sampled. For selenium, this does not always appear to be the case, but the 

potential effect of body size should be accounted for when collecting biota samples. Thus, 

body size (i.e., length and weight) should always be measured and data should be corrected 

for any relationship with size/weight (Tremblay et al. 1998, Wente 2004). 

Frequency of collection 

USEPA (2006) recommends that fish samples be collected at least once every two years to 

determine compliance with the fish tissue criterion for methylmercury. If a monitoring 

strategy for the North Bay is implemented, it may be appropriate to collect samples 

annually, at least initially. 

Compositing 

When collecting fish samples to support fish consumption advisories, USEPA (2000, 2006) 

recommends the collection of composite samples. What this means in that samples are 

collected from several fish individuals, but then the samples are homogenized and the 

laboratory analyzes only the single composite sample. This approach is designed to 
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minimize costs and to provide an indication of the average concentration. However, 

compositing biota samples makes evaluating relationships between 1) age, sex, size, time of 

year and 2) selenium concentration more difficult. Further, the costs for low level selenium 

analyses in tissues have fallen in recent years. Therefore, it is recommended that biota 

samples should not be composited, at least for organisms where it is possible to obtain 

samples large enough for analysis from a single individual (i.e., 2 g). 

Sample size 

If the Water Board elects to use the statistical techniques recommended above, then it is 

recommended that a minimum of eight samples be collected for a species during each 

sampling event. This will provide the necessary sample size to perform both background 

comparisons and the calculation of a UCL95. This is consistent with USEPA (2007b) 

recommendations. 

Fish tissue screening value protective of human health 

The State of California (OEHHA 2006) has calculated a screening value for selenium in fish 

tissue that is protective of fishermen (i.e., the human consumption fish). The screening value 

was calculated using toxicity data from USEPA (2007c). It should be noted that the 

calculations that the State of California (OEHHA 2006) used assume a 100% bioavailability 

of selenium in fish tissues to humans. However, as stated in the section on the toxicity of 

selenium to marine mammals, the bioavailability of selenium in animal tissues is generally 

lower (<25%) (Mahan 2001). This may need to be evaluated in more detail prior to 

implementing the fish tissue screening value protective of human health in the North Bay. 

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kestrels (Falco sparverius) fed selenomethionine and naturally incorporated selenium. 

Environmental Toxicology and Chemistry 17:2494–2497. 
North San Francisco Bay Selenium Toxicological Assessment April 2008 

Tetra Tech, Inc. A-1 

APPENDIX A 

WET WEIGHT TO DRY WEIGHT CONVERSIONS

Concentrations of selenium in tissues and animal feed are reported in either wet weight 

(ww) or dry weight (dw). Unfortunately, the two measurements are not equivalent. The 

selenium concentrations that were used in the analyses presented earlier in this report are 

presented in their original units in Table A-1. All of the selenium concentrations reported in 

Table A-1 were converted to dry weight using either the percent moisture reported in the 

original publication or the default percent moisture of 75% (USEPA 1993) for studies on 

fish where the percent moisture was not reported. The percent moisture used in the 

calculations is presented in Table A-1 and the dry weight selenium concentrations are 

reported in Table A-2. The formula used to covert wet weight to dry weight is as follows: 

100 percent moisture

100 Dry weight wet weight # % $

Additionally, not all studies measured selenium in all tissue types. USEPA (2004) provides 

equations to convert selenium concentrations in fish muscle and liver to whole body 

concentrations. The data that USEPA (2004) presented was also used to derive linear 

regression equations to convert selenium concentrations measured in one tissue type to 

another tissue type for all fish, except white sturgeon. The equations used are provided 

below and are given in order of preference used to estimate tissue concentrations. 

1. Selenium in fish muscles to whole body (USEPA 2004): 

( & & '') Sewholebody Semuscle % * $ ln8937.01331.0exp

2. Selenium in fish liver to whole body (USEPA 2004): 

& ' Sewholebody % #0 * 3071.02609. $ Seliver

3. Selenium in whole fish bodies to fish livers (p < 0.0001, r2

 = 0.73): 

( & & '') Seliver Sewholebody % * $ ln7628.06475.1exp

4. Selenium in whole fish bodies to fish muscle (p < 0.0001, r2

 = 0.96): 

( & & '') Semuscle Sewholebody % * $ ln031.10324.0exp

For white sturgeon, TetraTech obtained the raw data from the study conducted by Tashjian 

et al. (2006) and used that data to calculate the following regression equation (F1,22 = 222.42, 

Adjusted R2 = 0.91, p < 0.00001): 

( & & '') Sewholebody Semuscle % * $ ln6206.09586.0exp
April 2008 North San Francisco Bay Selenium Toxicological Assessment 

A-2 Tetra Tech, Inc. 

This regression equation was used to convert the muscle EC10 derived by Beckon to a 

whole body concentration as well as the muscle NOAEL and LOAEL from Linville (2006) 

to whole body concentrations. 

For the selenium toxicity studies on birds, only the concentration in the feed given to the 

birds was converted from wet weight to dry weight. That is presented in Table A-4 for 

toxicity studies and A-5 for production studies. 

References 

U.S. Environmental Protection Agency (USEPA). 1993. Wildlife Exposure Factors 

Handbook. Office of Research and Development. EPA/600/R-93/187. 
Table A-1

Fish Tissue Concentration Data

Measured Selenium Concentrations (mg/kg or ug/g)

Water Exposure Kidney Se (mg/kg) Liver Se (mg/kg) Muscle Se (mg/kg) Whole Body Se (mg/kg) Effect Used in

Fish Study Type Life stage Route NOAEL LOAEL NOAEL LOAEL NOAEL LOAEL NOAEL LOAEL dw/ww %moisture1 Notes Major Minor SMCV

Bluegill Cleveland (1993) fresh Juvenile Water - - - - - - 3.0 3.4 dw - 2 X

Bluegill Cleveland (1993) fresh Juvenile Diet - - - - - - 4.7 7.6 dw - 2 X X

Bluegill Coyle et al. (1993) fresh Adult Water/Diet - - - - - - 10 19 dw - 2,4 X X

Bluegill Coyle et al. (1993) fresh egg Water/Diet - - - - - - 23 41 dw - 2,4 X

Bluegill Doroshov et al. (1992) fresh Adult Diet - - 12.3 26.0 5.8 10.4 - - dw - - X X

Bluegill Finley (1985) fresh adult Diet - - 5.0 8.5 2.06 5.10 - - ww 75 - X

Bluegill Gillespie and Baumann (1986) fresh Fry/larvae Natural - - - - - - 3.1 28.2 dw - - X

Bluegill Gillespie and Baumann (1986) fresh Adult Natural - - - - - - 0.4 5.9 ww 75 - X

Bluegill Hermanutz et al. (1992) fresh adult Water/Diet - - 1.5 9.3 0.3 1.8 - - ww 75 - X X

Bluegill Lemly (1993a); Skorupa et al. 2004 fresh Juvenile Water/Diet - - - - - - 1.3 5.8 dw - - X X

Bluegill Sorensen et al. (1984) fresh Adult Natural - - 1.3 7.0 1.3 2.3 - - ww 75 - X

channel catfish Doroshov et al. (1992) fresh Adult Injection - - 7.3 12.5 - - - - dw - -

channel catfish Gatlin and Wilson (1984) fresh fingerling Diet - - - - 3.0 3.5 - - dw - - X X

chinook salmon Hamilton et al. (1990) fresh larvae Diet - - - - - - 5.4 10.8 dw - 6 X X

chinook salmon Hamilton et al. (1990) brackish juvenile Diet - - - - - - 12.6 23.2 dw - 6 X X

fathead minnow Dobbs et al. (1996) fresh larvae Water/Diet - - - - - - - 62 dw - - X X

fathead minnow Ogle and Knight (1989) fresh juvenile Diet - - - - 5.7 6.6 5.4 6.6 dw - - X X

fathead minnow Schultz and Hermanutz (1990) fresh egg Water/Diet - - - - - - 0.3 3.9 ww 80 - X X

green sunfish Lemly (1993b) fresh Adult Natural - - - - - - 12.4 14.7 dw - - X

rainbow trout Hicks et al (1984) fresh juvenile Diet 3.0 10.7 38.3 49.3 - - - - ww 75 - X X

rainbow trout Hilton and Hodson (1983) fresh juvenile Diet 0.7 3.8 0.6 21.0 - - 0.2 1.0 dw - 3, 4 X

rainbow trout Hilton and Hodson (1983) fresh juvenile Diet 3.8 7.7 21.0 71.7 - - 1.0 4.0 dw - 3, 4 X X

rainbow trout Hilton et al. (1980) fresh juvenile Water/Diet 9 37 42.0 95.0 - - - - dw - 2 X X

rainbow trout Holm et al. (2005) fresh adult Natural - - - - - 1.8 - - ww 75 - X

rainbow trout Holm et al. (2005) fresh egg Natural - - - - - - - 3.5 ww 61 - X

rainbow trout Hunn et al (1987) fresh Fry/larvae Water - - - - - - 0.5 0.9 dw - 4 X

rainbow trout Hunn et al (1987) fresh Fry/larvae Water - - - - - - 0.6 1 dw - 4 X

rainbow trout Vidal et al (2005) fresh Fry/larvae Diet - - - - - - 0.3 0.6 ww 75 - X X

Redear sunfish Sorensen et al. (1988) fresh Adult Natural 2.1 7.6 - - 0.3 5.0 - - ww 75 5 X

Sacramento splittail Teh et al. (2004) fresh juvenile Diet - - 23.0 26.8 10.1 15.1 - - dw - - X X

Striped Bass Coughlan and Velte (1989) fresh Adult Diet - - - - 1.1 3.5 - - ww 75 - X

white sturgeon Linville (2006) fresh Adult Diet - - 1.4 10.4 1.3 12.1 - - dw - - X X

white sturgeon Tashjian et al. (2006) fresh juvenile Diet 30.9 51.7 22.0 37.4 22.9 36.8 14.7 22.5 dw - - X X

Notes:

1 - a value of 75% is the default assumpttion (USEPA 19930

2 - concentration estimated from figures

3 - data from the low carbohydrate diet were not used as the authors reported that reduced weight of fish on this diet was likely due to food 

4 - dry weight assumed. Not explicilty stated whether dw or ww

5 - muscle concentration calculated using regression equation from Sorensen et al. (1982); i.e., muscle concentation = kidney concentration x 0.835 -1.375

6 - excludes treatment where fish were exposed to water from the San Luis Drain
Table A-2

Wet Weight to Dry Weight Conversion for Fish Tissue Concentration Data

Selenium Concentrations in Dry Weight (mg/kg-dw or ug/g-dw)

Water Exposure Kidney Se (mg/kg) Liver Se (mg/kg) Muscle Se (mg/kg) Whole Body Se (mg/kg) Effect Used in

Fish Study Type Life stage Route NOAEL LOAEL NOAEL LOAEL NOAEL LOAEL NOAEL LOAEL Major Minor SMCV

Bluegill Cleveland (1993) fresh Juvenile Water - - - - - - 3.0 3.4 X

Bluegill Cleveland (1993) fresh Juvenile Diet - - - - - - 4.7 7.6 X X

Bluegill Coyle et al. (1993) fresh Adult Water/Diet - - - - - - 10.0 19.0 X X

Bluegill Coyle et al. (1993) fresh egg Water/Diet - - - - - - 23.0 41.0 X

Bluegill Doroshov et al. (1992) fresh Adult Diet - - 12.3 26.0 5.8 10.4 - - X X

Bluegill Finley (1985) fresh adult Diet - - 20.0 33.9 8.2 20.4 - - X

Bluegill Gillespie and Baumann (1986) fresh Fry/larvae Natural - - - - - - 3.1 28.2 X

Bluegill Gillespie and Baumann (1986) fresh Adult Natural - - - - - - 1.5 23.6 X

Bluegill Hermanutz et al. (1992) fresh adult Water/Diet - - 6.0 37.2 1.2 7.2 - - X X

Bluegill Lemly (1993a); Skorupa et al. 2004 fresh Juvenile Water/Diet - - - - - - 1.3 5.8 X X

Bluegill Sorensen et al. (1984) fresh Adult Natural - - 5.2 28.0 5.2 9.2 - - X

channel catfish Doroshov et al. (1992) fresh Adult Injection - - 7.3 12.5 - - - - X X

channel catfish Gatlin and Wilson (1984) fresh fingerling Diet - - - - 3.0 3.5 - - X X

chinook salmon Hamilton et al. (1990) fresh larvae Diet - - - - - - 5.4 10.8 X X

chinook salmon Hamilton et al. (1990) brackish juvenile Diet - - - - - - 12.6 23.2 X X

fathead minnow Dobbs et al. (1996) fresh larvae Water/Diet - - - - - - - 62.0 X X

fathead minnow Ogle and Knight (1989) fresh juvenile Diet - - - - 5.7 6.6 5.4 6.6 X X

fathead minnow Schultz and Hermanutz (1990) fresh egg Water/Diet - - - - - - 1.6 19.6 X X

green sunfish Lemly (1993b) fresh Adult Natural - - - - - - 12.4 14.7 X

rainbow trout Hicks et al (1984) fresh juvenile Diet 12.0 42.8 153.2 197.2 - - - - X X

rainbow trout Hilton and Hodson (1983) fresh juvenile Diet 0.7 3.8 0.6 21.0 - - 0.2 1.0 X

rainbow trout Hilton and Hodson (1983) fresh juvenile Diet 3.8 7.7 21.0 71.7 - - 1.0 4.0 X X

rainbow trout Hilton et al. (1980) fresh juvenile Water/Diet 9.0 37.0 42.0 95.0 - - - - X X

rainbow trout Holm et al. (2005) fresh adult Natural - - - - - 7.2 - - X

rainbow trout Holm et al. (2005) fresh egg Natural - - - - - - - 9.0 X

rainbow trout Hunn et al (1987) fresh Fry/larvae Water - - - - - - 0.5 0.9 X

rainbow trout Hunn et al (1987) fresh Fry/larvae Water - - - - - - 0.6 1.1 X

rainbow trout Vidal et al (2005) fresh Fry/larvae Diet - - - - - - 1.2 2.3 X X

Redear sunfish Sorensen et al. (1988) fresh Adult Natural 8.2 30.5 - - 1.3 20.0 - - X

Sacramento splittail Teh et al. (2004) fresh juvenile Diet - - 23.0 26.8 10.1 15.1 - - X X

Striped Bass Coughlan and Velte (1989) fresh Adult Diet - - - - 4.4 14.0 - - X

white sturgeon Linville (2006) fresh Adult Diet - - 1.4 10.4 1.3 12.1 - - X X

white sturgeon Tashjian et al. (2006) fresh juvenile Diet 30.9 51.7 22.0 37.4 22.9 36.8 14.7 22.5 X X
Table A-3

Conversion of Fish Tissue Concentration Data Among Tissue Types

Selenium Concentrations in Dry Weight (mg/kg-dw or ug/g-dw)

Water Exposure Kidney Se (mg/kg) Liver Se (mg/kg) Muscle Se (mg/kg) Whole Body Se (mg/kg) Effect Used in

Fish Study Type Life stage Route NOAEL LOAEL NOAEL LOAEL NOAEL LOAEL NOAEL LOAEL Major Minor SMCV

Bluegill Cleveland (1993) fresh Juvenile Water - - 12.0 13.2 3.2 3.6 3.0 3.4 X

Bluegill Cleveland (1993) fresh Juvenile Diet - - 16.8 24.5 5.1 8.4 4.7 7.6 X X

Bluegill Coyle et al. (1993) fresh Adult Water/Diet - - 30.1 49.1 11.1 21.5 10.0 19.0 X X

Bluegill Coyle et al. (1993) fresh egg Water/Diet - - - - - - 23.0 41.0 X

Bluegill Doroshov et al. (1992) fresh Adult Diet - - 12.3 26.0 5.8 10.4 5.5 9.3 X X

Bluegill Finley (1985) fresh adult Diet - - 20.0 33.9 8.2 20.4 7.5 16.9 X

Bluegill Gillespie and Baumann (1986) fresh Fry/larvae Natural - - 12.3 66.3 3.3 32.3 3.1 28.2 X

Bluegill Gillespie and Baumann (1986) fresh Adult Natural - - 7.0 58.0 1.5 26.9 1.5 23.6 X

Bluegill Hermanutz et al. (1992) fresh adult Water/Diet - - 6.0 37.2 1.2 7.2 1.3 6.7 X X

Bluegill Lemly (1993a); Skorupa et al. 2004 fresh Juvenile Water/Diet - - 6.3 19.9 1.4 6.3 1.3 5.8 X X

Bluegill Sorensen et al. (1984) fresh Adult Natural - - 5.2 28.0 5.2 9.2 5.0 8.3 X

channel catfish Doroshov et al. (1992) fresh Adult Injection - - 7.3 12.5 2.1 3.8 2.0 3.6 X X

channel catfish Gatlin and Wilson (1984) fresh fingerling Diet - - - - 3.0 3.5 3.1 3.5 X X

chinook salmon Hamilton et al. (1990) fresh larvae Diet - - 18.8 31.9 5.9 12.0 5.4 10.8 X X

chinook salmon Hamilton et al. (1990) brackish juvenile Diet - - 35.9 57.2 14.1 26.4 12.6 23.2 X X

fathead minnow Dobbs et al. (1996) fresh larvae Water/Diet - - - 121.0 - 72.8 - 62.0 X X

fathead minnow Ogle and Knight (1989) fresh juvenile Diet - - 18.8 21.9 5.7 6.6 5.4 6.6 X X

fathead minnow Schultz and Hermanutz (1990) fresh egg Water/Diet - - - - - - 1.6 19.6 X X

green sunfish Lemly (1993b) fresh Adult Natural - - 35.4 40.3 13.8 16.5 12.4 14.7 X

rainbow trout Hicks et al (1984) fresh juvenile Diet 12.0 42.8 153.2 197.2 - - 46.8 60.3 X X

rainbow trout Hilton and Hodson (1983) fresh juvenile Diet 0.7 3.8 0.6 21.0 0.2 1.0 0.2 1.0 X

rainbow trout Hilton and Hodson (1983) fresh juvenile Diet 3.8 7.7 21.0 71.7 1.0 4.3 1.0 4.0 X X

rainbow trout Hilton et al. (1980) fresh juvenile Water/Diet 9.0 37.0 42.0 95.0 - - 12.6 28.9 X X

rainbow trout Holm et al. (2005) fresh adult Natural - - - - - 7.2 - 6.7 X

rainbow trout Holm et al. (2005) fresh egg Natural - - - - - - - 9.0 X

rainbow trout Hunn et al (1987) fresh Fry/larvae Water - - 3.2 4.7 0.5 0.9 0.5 0.9 X

rainbow trout Hunn et al (1987) fresh Fry/larvae Water - - 3.7 5.5 0.7 1.1 0.6 1.1 X

rainbow trout Vidal et al (2005) fresh Fry/larvae Diet - - 6.1 9.9 1.3 2.5 1.2 2.3 X X

Redear sunfish Sorensen et al. (1988) fresh Adult Natural 8.2 30.5 - - 1.3 20.0 1.5 16.6 X

Sacramento splittail Teh et al. (2004) fresh juvenile Diet - - 23.0 26.8 10.1 15.1 9.0 12.9 X X

Striped Bass Coughlan and Velte (1989) fresh Adult Diet - - - - 4.4 14.0 4.3 12.1 X

white sturgeon Linville (2006) fresh Adult Diet - - 1.4 10.4 1.3 12.1 3.1 12.3 X X

white sturgeon Tashjian et al. (2006) fresh juvenile Diet 30.9 51.7 22.0 37.4 22.9 36.8 14.7 22.5 X X

Note: shaded cells represent predicted concentrations. See text for explanation.
Table A-4

Wet Weight to Dry Weight Conversion for Bird Dietary Concentration Data

Measured Selenium Concentrations (mg/kg or ug/g)

w

Measured Diet Se (mg/kg) Diet Se (mg/kg-dw) Effect

Bird Study Life stage Chemical form NOAEL LOAEL dw/w %moisture NOAEL LOAEL Major Minor

Chicken Elzubier and Davis (1988) Chick selenite 0.15 10 ? ? 0.15 10 X

Chicken (Hubbard) Jensen (1976) Chick selenite 0.9 4.3 dw - 0.9 4.3 X

Chicken (leghorn) Jensen (1976) Chick selenite 4.3 13.5 dw - 4.3 13.5 X

Chicken Lowry and Baker (1989) Chick selenite; SeMet 0 15 ? ? 0 15 X

Chicken Ort and Latshaw (1978) Adult selenite 3 5 ? ? 3 5 X

Common eider Franson et al. (2007) Adult SeMet 20.6 57.7 dw - 20.6 57.7 X

Mallard Albers et al. (1996) Adult SeMet 20 40 ww 10.5 22.3 44.7 X

Mallard Green and Albers 1996 Adult SeMet 10 20 ww 10.5 11.2 22.3 X

Mallard Heinz and Hoffman, 1998 Adult SeMet 0.4 8.8 ww 10 0.4 9.8 X

Mallard Heinz et al., 1987 Adult selenite 10 25 ww 8.5 10.9 27.3 X

Mallard Heinz et al., 1987 Adult SeMet 0 10 ww 8.5 0.0 10.9 X

Mallard Heinz et al., 1988 Duckling selenite 10.1 20.1 ww 7 10.9 21.6 X

Mallard Heinz et al., 1988 Duckling SeMet 10.1 20.1 ww 7 10.9 21.6 X

Mallard Heinz et al., 1989 Adult SeMet 4.1 8.1 ww 10 4.6 9.0 X

Mallard Heinz et al., 1989 Adult SeCysteine 10.1 - ww 10 11.2 - X

Mallard Heinz et al. (1996) Duckling SeMet 15.34 30.34 ww1 10 17.0 33.7 X

Mallard Hoffman and Heinz (1998) Adult SeMet 0.4 10.4 ww 10 0.4 11.6 X

Mallard Hoffman et al. (1991) Duckling SeMet; low protein 0.2 15.2 ww 10 0.2 16.9 X

Mallard Hoffman et al. (1991) Duckling SeMet 15.2 60.2 ww 10 16.9 66.9 X

Mallard Hoffman et al. (1992a) Duckling SeMet 0.2 15.2 ww 10 0.2 16.9 X

Mallard Hoffman et al., 1992b Duckling SeMet 0.2 15.2 ww 8 0.2 16.5 X

Mallard Hoffman et al., 1992b Duckling SeMet 15.2 60.2 ww 8 16.5 65.4 X

Mallard Hoffman et al., 1992b Duckling SeMet; low protein 15.2 60.2 ww 8 16.5 65.4 X

Mallard Hoffman et al., 1992b Duckling SeMet; low protein 0.2 15.2 ww 8 0.2 16.5 X

Mallard Hoffman et al., 1992b Duckling SeMet; high protein 0.2 15.2 ww 8 0.2 16.5 X

Mallard Hoffman et al. (1996) Duckling SeMet, Se yeast, Se in wheat 0.2 15.2 ww1 10 0.2 16.9 X

Mallard O'Toole and Raisbeck (1997) Flightling SeMet 10.32 25.32 ww 25 13.8 33.8 X

Mallard Stanley et al. (1994) Adult SeMet 0.37 6.5 dw - 0.37 6.5 X

Mallard Stanley et al. (1996) Adult SeMet 3.5 7 ww 10 3.9 7.8 X

Pheasant Latshaw et al. (2004) Adult unknown - 9.3 dw - - 9.3 X

Screech owl Wiemeyer and Hoffman (1996) Adult SeMet 8.8 30 dw - 8.8 30 X

Screech owl Wiemeyer and Hoffman (1996) Adult SeMet 0.3 8.8 dw - 0.3 8.8 X

Notes: 1 - a %moisture of 10 was assumed based on other studies by the same authors where a %moisture was specified.
Table A-5

Summary of Chicken Enhanced Production Studies

Measured Selenium Concentrations (mg/kg or ug/g)

Measured NOAEL NOAEL Effects

Bird Study Life stage Dietary form (mg/kg) dw/ww %moisture (mg/kg-dw) Survival Weight Egg output Reproductive success

Chicken Payne et al. (2005) Adult selenite 2.6 ? - 2.6 X

Chicken Payne et al. (2005) Adult Se-yeast 2.9 ? - 2.9 X

Chicken Ryu et al. (2005) 1 day selenite 8.2 dw - 8.2 X

Japanese quail Biswas et al. (2006) 1 day selenite 1.2 ? - 1.2 X X

Japanese quail Sahin and Kucuk (2001) 10 day selenite 0.2 ww 10.32 0.2 X

North San Francisco Bay Selenium Toxicological Assessment April 2008 

Tetra Tech, Inc. B-1 

APPENDIX B 

THRESHOLD VALUES FOR SELENIUM IN GREAT SALT LAKE

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North San Francisco Bay Selenium Toxicological Assessment April 2008 

Tetra Tech, Inc. C-1 

APPENDIX C 

ALLOMETRIC ADJUSTMENT OF DIETARY CONCENTRATIONS

The Appendix explains the approach used to allometrically adjust dietary toxicity threshold 

for birds. In ecological risk assessments for bids, dose-based toxicity reference values 

(TRVs) are used to estimate toxic effects. Toxicity reference values are calculated using the 

following equation: 

BW

AFFCIREPC TRVt

$ $ $ %

where: 

TRVt = Toxicity reference value for the test species. 

EPC = Exposure point concentration; i.e., the dietary concentration of a 

chemical that produced the toxic effects measured in the toxicity 

study (mg/kg-dw). 

IR = Food ingestion rate (mg/day-dw). 

FC = Fraction contacted (unitless; assumed to be 1). 

AF = Assimilation fraction (unitless; assumed to be 1). 

BW = Body weight (kg). 

Toxicity reference values are allometrically adjusted using the following equation (Sample 

and Arenal 1999): 

)b1(

a

a t BW

BW TRV TRV

#

+

+

,

-

.

.

/

0

$% t

where: 

TRVa = Allometrically adjusted toxicity reference value. 

TRVt = Toxicity reference value for the test species. 

BWt = Body weight of the test species. 

BWa = Body weight of species for which the allometric adjustment is 

being performed. 

b = Allometric scaling factor (unitless). 
April 2008 North San Francisco Bay Selenium Toxicological Assessment 

C-2 Tetra Tech, Inc. 

The 

)b1(

BWa

BW #

+

+

,

-

.

.

/

0 t term in the equation above is also referred to as an “allometric adjustment 

factor”. Combining the above two equations, and eliminating the variables assumed to equal 

1, produces the following equation: 

)b1(

a

a t

BW

BW

BW

IREPC

BW

IREPC #

+

+

,

-

.

.

/

0 1 $

2

3 4

5

6 $ % $ t

t

t

a

a

By re-arranging the equation above, it is possible to derive an allometrically adjusted dietary 

concentration protective of specific toxic effect. The re-arranged equation is as follows: 

)b1(

a

t

a

a

a BW

BW

BW

IREPC

IR

BW EPC

#

+

+

,

-

.

.

/

0 1 $

2

3 4

5

6 $

1 $

2

3 4

5

6 % t

t

t

Allometrically adjusted dietary thresholds were calculated for the avian species of concern 

in the Bay identified by USFWS (2007): 

! Bald eagle 

! California clapper rail 

! Greater scaup 

! Lesser scaup 

! White-winged scoter 

! Surf scoter 

! Black scoter 

The test species from which the TRVs were allometrically adjusted include: 

! Leghorn chicken 

! Mallard 

! Screech owl 

The body weights for all of these birds are provided in Table C-1 below. Ingestion rates for 

each bird species were calculated from their body weights using the equations provided by 

Nagy (2001). The ingestion rate equations and calculated ingestion rates are also shown in 

Table C-1 below. 
North San Francisco Bay Selenium Toxicological Assessment April 2008 

Tetra Tech, Inc. C-3 

Table C-1 

Body weights and ingestion rates. 

Bird species 

Body Weight 

(g) Ingestion Rate Equationf Ingestion Rate (mg/day-dw) 

Bald eaglea 5,275 Carnivorous birds y = 0.849*(x)^0.663 249,311 

California clapper raila 346 Marine birds y = 0.880*(x)^0.658 41,228 

Greater scaupa 959 Marine birds y = 0.880*(x)^0.658 80,634 

Lesser scaupa 663 Marine birds y = 0.880*(x)^0.658 63,247 

White-winged scotera 1,917 Marine birds y = 0.880*(x)^0.658 127,189 

Surf scoterb 1,047 Marine birds y = 0.880*(x)^0.658 85,429 

Black scotera 987 Marine birds y = 0.880*(x)^0.658 82,198 

Mallardc 1,214 Marine birds y = 0.880*(x)^0.658 94,167 

Chickend 1,800 Galliformes y = 0.088*(x)^0.891 69,973 

Screech Owle 186 Carnivorous birds y = 0.849*(x)^0.663 27,139 

Notes: 

a - female body weight from USFWS (2007) 

b - female body weight from BNA (2007) 

c - body weight is from Heinz et al. (1989) and is average of females at time of sacrifice in all treatments except 16 

mg/kg 

d - egg laying leghorn chickens average 1,800 g (National Research Council 1994) 

e - body weight is from Wiemeyer and Hoffman (1996) and is average of females at time of sacrifice in all treatments 

f - from Nagy (2001) 

The body weights and ingestion rates from Table C-1were used to calculate allometric 

adjustment factors and allometrically adjusted dietary threshold concentrations for the 

species of concern in the Bay (Table C-2). 

Table C-2 

Allometric adjust factors and allometrically adjusted dietary thresholds. 

Allometric Adjustment Factor (-) 

for Dietary Screening Values from 

Dietary Screening 

Value (mg/kg-dw) Adjusted from 

Bird species Mallard Chicken Screech Owl Mallarda

 Chickenb Screech Owlc

Bald eagle 1.34 1.24 1.95 9.7 4.0 104.3 

California clapper rail 0.78 0.72 - 2.2 0.9 - 

Greater scaup 0.95 0.88 - 3.9 1.6 - 

Lesser scaup 0.89 0.82 - 3.2 1.3 - 

White-winged scoter 1.10 1.01 - 5.6 2.3 - 

Surf scoter 0.97 0.90 - 4.1 1.7 - 

Black scoter 0.96 0.89 - 3.9 1.6 - 

Notes: 

a - EC10 for reduced hatching success from Adams et al. (2003) Ohlendorf (2007) of 4.4 mg/kg-dw 

b - effect threshold for reduced hatching success of 3.9 mg/kg-dw from Ort and Latshaw (1978) 

c - effect threshold for reduced hatching success of 17.3 mg/kg-dw from Wiemeyer and Hoffman (1996) 

References 

Adams WJ, Brix KV, Edwards M, Tear LM, Deforest DK, Fairbrother A. 2003. Analysis of 

field and laboratory data to derive selenium toxicity thresholds for birds. Environmental 

Toxicology and Chemistry 22:2020–2029. 

Birds of North America Online. 2007. Cornell Laboratory of Ornithology, Ithaca, NY. 

Available online at: http://bna.birds.cornell.edu/BNA/. 

Heinz GH, Hoffman DJ, Gold LG. 1989. Impaired reproduction of mallards fed an organic 

form of selenium. Journal of Wildlife Management 53:418–428. 
April 2008 North San Francisco Bay Selenium Toxicological Assessment 

C-4 Tetra Tech, Inc. 

Nagy KA. 2001. Food requirements of wild animals: Predictive equations for free-living 

mammals, reptiles, and birds. Nutrition Abstracts and Reviews, Series B: Livestock Feeds 

and Feeding 71:2R-12R. 

National Research Council. 1994. Nutrient Requirements of Poultry. Ninth Edition. National 

Academy Press, Washington, D.C. 

Ohlendorf HM. 2007. Threshold Values for Selenium in Great Salt Lake: Selections by the 

Science Panel. Submitted to Great Salt Lake Science Panel. February 28, 2007. 

Ort JF and Latshaw JD. 1978. The toxic level of sodium selenite in the diet of laying 

chickens. Journal of Nutrition 108:1114-1120. 

Sample BE and Arenal CA. 1999. Allometric models for interspecies extrapolation of 

wildlife toxicity data. Bulletin of Environmental Contamination and Toxicology 62:653-663. 

USFWS (U.S. Fish and Wildlife Service). 2007. Species at Risk from Selenium Exposure in 

the San Francisco Estuary. 

Wiemeyer SN and Hoffman DJ. 1996. Reproduction in eastern screech-owls fed selenium. 

Journal of Wildlife Management 60:332-341.