Technica

Spe

U. S.

F

D

Env

al Rev

ecific S

 DEPARTM

FISH AND W

Division of E

vironmental 

iew: S

Seleniu

Jan

ENT OF TH

WILDLIFE S

nvironmenta

Contaminan

moky 

um Cri

nuary 2012

HE INTERIO

ERVICE 

al Quality 

nts Program

Canyo

iterion

2

R

on Min

 Repo

e Sitert 
Table of Contents 

Title Page............................................................................................. 1 

Table of Contents................................................................................. 2 

Executive Summary............................................................................. 3 

Reviewer’s Preface.............................................................................. 4 

SECTION I: Scientific Issues Warranting Reconsideration 

1.0 Introduction.................................................................................... 5 

2.0 Site Setting.................................................................................... 5 

3.0 Literature Review.......................................................................... 6 

4.0 Field Monitoring – Exposure Characteristics................................. 9 

5.0 Laboratory Studies – Effects Characteristics................................. 9 

6.0 Unaddressed Issues......................................................................11 

SECTION II: Recommended Recalculations 

(1) Hatch to 15-days Post Swim-up Fry Survival Endpoint................13 

(2) Larval Deformity Endpoint............................................................16 

(3) Regressing X on Y for Estimation of Critical Whole Body 

 Tissue Concentrations..................................................................19 

(4) Test of Assumption that the 14 ppm Proposed Critical 

 Whole Body Tissue Concentration Derived from Brown 

 Trout Tissue Partitioning Data Would Generally Keep 

 Egg Selenium Values below a Critical Value of 21.63 ppm 

 Across a Multi-species Community of Fish..................................22 

Reviewer’s Summary..........................................................................23 

Literature Cited....................................................................................25 

Appendix 1: Peer Reviews and Peer Reviewer Curriculum Vitae.......27 
U.S. Fish & Wildlife Service Technical Review: Smoky Canyon 

Mine Site-Specific Selenium Criterion Report (August 2010 

Draft) 

Executive Summary: 

In response to a formal Congressional request, the Division of Environmental Quality 

(DEQ) of the U.S. Fish and Wildlife Service (FWS) prepared a scientific technical review 

of the draft report titled, “Interpretive Findings for Field and Laboratory Studies and 

Literature Review in Support of a Site-specific Selenium Criterion, Smoky Canyon 

Mine.” The base report and numerous appendices, in sum totaling more than 1,200 pages 

of documentation, were closely examined. 

Multiple distinct scientific issues warranting reconsideration were identified. Five of the 

most substantive issues included: (1) The lack of valid field controls for interpretation of 

subsequent toxicity testing as indicated by the data collected from upstream sampling 

sites versus a true spatial reference site; (2) The failure to properly incorporate complete 

absence of “swim-up” among some experimental groups into exposure-response toxicity 

modeling; (3) Systematically biased low and environmentally unrealistic quantification of 

larval deformity rates; (4) Unjustified extrapolation of the brown trout ratio for 

partitioning of selenium between egg tissue and whole body tissue to all other cooccurring species of fish; and (5) The absence of any wildlife risk analysis despite the 

Clean Water Act’s explicit mandate for protection of fish, shellfish, and wildlife. 

An attempt to make reasonable data corrections and recalculate the exposure-response 

relationship for the endpoint of “15-day post swim-up survivorship” resulted in a set of 

data unsuitable for estimating an EC-10 or EC-20 because of a lack of intermediateeffects response outcomes. An attempt to make reasonable data corrections and 

recalculate the exposure-response relationship for the larval deformity endpoint produced 

EC-10 and EC-20 point estimates that were substantively lower than those originally 

produced in the draft report. The recalculations for the larval deformity endpoint could 

be improved by further resolving two key uncertainties: (1) the use of hatchery control 

data as a surrogate for valid field control data; and (2) the use of confounded data 

summaries for incidences of larval deformities instead of the unconfounded raw data 

(which J.R. Simplot Co. declined to provide for this review). An examination of 

selenium tissue partitioning in brown trout relative to a 12-species community of 

Colorado River fishes suggested that in the majority of fish species their egg selenium 

concentrations would exceed the proposed criterion if the whole-body selenium threshold 

for brown trout were applied to all other species. 

Three external, independent, peer reviewers concurred with the DEQ/FWS technical 

critiques in 94.8% (N=58) of their point-by-point peer review evaluations. 
Selected Technical Review Comments for: 

Interpretive Findings for Field and Laboratory Studies and 

Literature Review in Support of a Site-Specific Selenium 

Criterion, Smoky Canyon Mine (August 2010) 

REVIEWER’S PREFACE

Via a letter dated March, 30, 2011, from U.S. Senator Barbara Boxer to U.S. Fish and 

Wildlife Service (FWS) Acting Director Rowan Gould, FWS received a technical 

assistance request. The request, on behalf of the Senate Committee on Environment and 

Public Works (EPW), was for a scientific technical review of a draft report (Report) 

titled, “Interpretive Findings for Field and Laboratory Studies and Literature Review in 

Support of a Site-Specific Selenium Criterion, Smoky Canyon Mine.” The Report was 

prepared for the J.R. Simplot Company by two consulting firms and the focus of the 

Report is J.R. Simplot’s Smoky Canyon phosphate mine. As part of earlier investigations 

resulting from an Administrative Order on Consent (AOC) entered into by Simplot, the 

Idaho Department of Environmental Quality (IDEQ), the U.S. Forest Service (USFS) and 

the U.S. Environmental Protection Agency (USEPA) together forming the Site Specific 

Selenium Criterion (SSSC) Workgroup, it was determined that disposal areas for mining 

waste rock were the source of selenium pollution to down gradient streams in excess of 

current water quality standards. Source control actions have been identified and 

implemented, but will not immediately reduce selenium discharges from a contaminated 

aquifer. Thus, in the interim, the possibility of relaxing the current water standard on a 

site-specific basis is being pursued. The Report provides part of the technical basis that 

would be necessary for a site-specific variance from the currently applicable water 

quality standards for selenium. 

FWS staff scientists have been continuously involved with the scientific study of the 

ecotoxicology of selenium since severe toxic impacts on fish and wildlife from selenium 

pollution emerged as a major issue at Kesterson National Wildlife Refuge (CA) in the 

mid-1980s. In addition to conducting original research on the ecotoxicology of selenium, 

FWS scientists are routinely asked to serve as participants on science panels (e.g., 

Selenium Science Panel for a site-specific selenium criterion for the Great Salt Lake, 

Utah) and to serve as scientific peer reviewers for prominent professional journals such 

as Science, Environmental Science and Technology, and many others. The present 

review was conducted as a general scientific peer review. In other words, as if the Report 

had been submitted for publication and a journal editor had requested a peer review. The 

only exception to that approach is that the time frame for this review was limited relative 

to the length and complexity of the report and appendices (greater than 1,200 pages), thus 

only a selected review of prioritized topics was possible. The following review 

comments follow in the order of the outline of the Report beginning with the Report 

Introduction. Bold face type was used to identify language extracted directly from the 

Report. 
SECTION I: Scientific Issues Warranting Reconsideration 

1.0 INTRODUCTION

 

Interpretive findings for Yellowstone cutthroat trout (YCT) are brought into question by 

the disclaimer presented in this section stating that this “... Interpretive Report precedes 

the completion of the YCT studies reporting...” If the YCT toxicity studies are not yet 

completed, then, by normal scientific protocol, it would be premature to attempt to 

interpret those studies. [p. 3]

The Report exhibits some confusion regarding the basis for EPA’s current national water 

quality chronic criterion for selenium of 5 parts per billion (ug/L). It states that it is 

“...based on aqueous exposure only and includes no dietary component of exposure...” 

which is erroneous. The Report authors correctly state that the criterion is based on field 

data from Belews Lake in North Carolina, but they fail to link that with the fact that fish 

at Belews Lake experienced both aqueous and dietary exposure to selenium as is typical 

of a field study at any selenium-polluted lake. [p. 5] 

The Report again exhibits a degree of confusion with regard to the basis for EPA’s draft 

chronic criterion revision of 2004 stating that it also was based on data from Belew’s 

Lake in North Carolina. It was not; it was based on a laboratory experiment (Lemly 

1993). To add to this confusion, the authors then ask the reader to compare GMCVs 

(genus mean chronic values) for trout from EPA’s 2004 draft document to the lower 

bluegill-based draft criterion proposal while failing to mention that the bluegill result is 

for an entirely different, and more sensitive, toxicity test (multiple stressors of Se and 

temperature) than the trout GMCVs (single stressor, Se only). Thus, the implication that 

trout are more tolerant than bluegill is based on a false comparison. [p. 6]

It is correctly noted in the Report that USEPA commissioned a follow-up study meant to 

replicate and extend the work of Lemly (1993), but the Report fails to note that the 

follow-up study did not replicate Lemly’s work. Lemly’s work focused on the interactive 

effects of selenium exposure and winter-stress syndrome (a physiological response to 

declining water temperature and decreasing photo period expressed as lipid loss). The 

follow-up study did not decrease photo period and only lowered water temperature, 

which failed to induce lipid loss (i.e., failed to induce winter-stress syndrome). Because 

the follow-up study did not mimic “winter”, unlike Lemly’s work, it does not refute or 

replace Lemly’s original work; furthermore this limited study is of unknown applicability 

to free-ranging fish. [p. 6]

2.0 SITE SETTING

The field sampling design is predicated on an assumption that sampling locations 

upstream of selenium discharges to Crow Creek will represent water quality conditions 
“...unimpacted by mining activities.” However, as will be discussed below, the 

biological data contradict that assumption, which has important implications for the 

laboratory trout toxicity studies (also discussed below). [p. 11] 

3.0 LITERATURE REVIEW

The literature review is uneven. For example, Besser et al.’s (2008) reporting of a low 

toxicity threshold for desert pupfish is dismissed based on an unpublished memo (Canton 

and DeForest 2009) prepared on behalf of the Mining Association and Rio Tinto (a 

mining corporation); a memo that is not easily accessible to the scientific community, as 

no URL link to a public posting of this privately held memo is provided in the References 

section of the report [p. 22 and p. 74]. Yet, results that support a high toxicity threshold 

reported by Kennedy et al. (2000) are accepted uncritically with no mention of a peerreviewed, published critique of Kennedy et al. (2000) by Hamilton and Palace (2001) [p. 

24 and p. 25]. Hamilton and Palace argued, for example, that the study was 

compromised by unexplained high mortality in eggs obtained from the reference site and 

by such a narrow exposure interval among most non-reference eggs that there would be 

“... little hope of deciphering [statistically] significant differences in reproductive 

success.” Issues also could be taken with many other specifics of the literature review, 

but ultimately the Report’s derivation of a proposed site-specific selenium criterion is not 

based substantively on the literature review, so the limited time available for this review 

will be allocated to higher priority comments.

4.0 FIELD MONITORING – EXPOSURE CHARACTERISTICS

It is stated in the Report that, “Sulfate has been shown to potentially reduce selenium 

bioaccumulation and resulting toxicity.” Presumably this is deemed significant because 

sulfate concentrations were found to positively co-vary with selenium concentrations in 

the study area and national selenium standards are not based on studies from high sulfate 

waters. However, it is not noted in the Report that sulfate inhibition of selenate uptake, 

while demonstrated in the laboratory for environmentally unrealistic conditions (very 

short-term exposures without other chemical species of selenium present), has never been 

demonstrated to meaningfully alter bioaccumulation of selenium into aquatic organisms 

under field conditions, even in selenate-dominated waters (e.g., Birkner 1978). The 

authors of lab studies demonstrating sulfate reduction of selenium uptake in aquatic 

organisms clearly recognized the strict limits of their work for extrapolation to natural 

waters. For example Hansen et al. (1993:77) wrote, “Thus, at this time, it does not 

appear that we have sufficient evidence to justify the consideration of sulfate as a factor 

in the regulation of Se in aquatic environments.” Williams et al. (1994:452) wrote, “At 

present there is little information available that allows us to assess how relevant this 

study’s conclusions will be in natural waters containing a complex assemblage of 

selenium [chemical] species.” Finally, Ogle and Knight (1996:278) reported that for 

water concentrations of selenium near the national criterion of 5 ppb, “... the differences 

[in selenium bioaccumulation and toxicity] between extremely different sulfate 
concentrations are not significant...” In short, there is not a reasonable scientific basis to 

expect the sulfate concentrations in the study area to make any difference with regard to a 

site-specific selenium criterion. [p. 33] 

Because of the hydrological connectivity between the upstream putative background 

sampling sites and the selenium source waters the SSSC Workgroup collectively agreed 

that a reference site completely outside the phosphate mining district was also needed. 

The chosen reference site was on the South Fork Tincup Creek [p. 14]. The Report later 

states, “With few exceptions, concentrations of selenium in both trout and sculpins 

tissue from upstream Crow Creek and Deer Creek locations ( [putative] background) 

are not significantly different from the reference location tissue concentrations.” [p. 

34]. 

However, that statement is contradicted by data presented in Figure 4-8, Table 5-2, 

Appendix A: Table 4 and the discussion of statistical results on page 6 of Appendix A. 

Statistical test results presented in Appendix A: Table 4, show that trout whole body 

tissue selenium at the reference site is statistically significantly lower than at the 

upstream purported background sampling sites on Crow Creek and Deer Creek. This is 

also visually evident in Figure 4-8. Table 5-2 reports a trout tissue selenium value of 

2.56 ug/g dw at the regionally appropriate reference site outside the mining district, 

which is typical for fish from waters unpolluted with selenium (see Walsh et al. 1977; 

Lowe et al. 1985 and summary in Skorupa 1998: Table 1). By comparison, fish tissue 

values in upstream Crow Creek and Deer Creek, the purported background streams, 

averaged from about 6-9 ug/g dw which is well above the unpolluted fish tissue norms 

established by national monitoring programs. Trout movement up and downstream into 

and out of the exposed stream segments and the purported background segments 

constitute the most plausible explanation for these data. Accordingly, the assumption that 

the upstream sampling sites represent background conditions uninfluenced by mining 

seems unlikely to be valid, especially given the absence of barriers to fish movement in 

the study system. 

It should be noted in the Report that brown trout are known to undergo vitellogenesis, the 

stage of egg formation when selenium is incorporated into fish eggs, for 6-7 months prior 

to spawning (Estay et al. 2003). This extended period allows ample time for long 

distance movement of fish up and downstream. This interpretation of the fish tissue data 

is further supported by the results illustrated for periphyton in Figure 4-14. Periphyton is 

capable of only passive movement with water currents and therefore cannot move 

upstream. In contrast to the mobile trout, the tissue selenium of periphyton is not

significantly different (see also Appendix A, Table 14) between the regionally 

appropriate reference site on South Fork Tincup Creek and the upstream Crow Creek and 

Deer Creek sites (see also Appendix A, Table 14). 

In summary, Figure 4-1 shows tight concordance of water selenium concentrations (< 1-2 

ppb) for the reference site and the purported background upstream sites, Figure 4-14 

shows tight concordance of periphyton selenium concentrations (2-3 ppm) for the 

reference site and the upstream sites, yet the trout selenium concentrations are distinctly 
discordant at the reference site versus the purported background upstream sites. The 

much higher trout selenium at the upstream sites relative to the regionally appropriate 

reference site would be the unexpected result, not the low reference site trout selenium, 

according to current models of selenium transfer from a periphyton food web base up to 

fish (see Presser and Luoma 2010). The fact that the fish sampled at the purported 

background locations appear to have been influenced by mining has important 

implications for the brown trout toxicity study in light of the fact that no spawning brown 

trout that could have served as valid field controls were collected from South Fork 

Tincup Creek (discussed further below). 

The Report indirectly highlights the fundamental problem with population level 

evaluations of trout when it is stated, For this study, trout populations are being 

evaluated to assess if there are any obvious [underlining mine] negative impacts to 

those fish populations due to selenium.” [p. 40] 

Clearly, “...obvious negative impacts..” are rarely evident at the threshold levels of 

toxicity that are the appropriate focus for establishing water quality criteria; especially in 

demographically open populations that can be maintained by immigration. Using fish 

density measurements or age-class profiles as the baseline criteria to establish an effect is 

indefensible due to the high baseline variability of such coarse metrics (see review of this 

topic in Janz et al. 2010: Section 6.7.3). By relying on such coarse metrics the Report 

fails to clearly distinguish between measures of habitat attractiveness and habitat quality. 

What these investigators should be looking at in the field are specific demographic 

parameters that would allow demographic modeling. For example, Vasterling (2003) 

took that approach to studying riparian birds in the Idaho phosphate district and found 

that even though the species she studied were highly abundant, they were not selfreplacing and that their persistence (at least during the study years) depended on 

immigration. She also found evidence for localized selenium toxicity, despite 

documenting abundant populations. 

Low quality polluted habitats that poison fish and birds can, nonetheless, be highly 

attractive. For example, Kesterson Reservoir in California was a site that caused 

catastrophic levels of reproductive toxicity among birds including complete (100%) 

reproductive failure, yet year after year very high densities of breeding water birds could 

be found at Kesterson Reservoir, with these demographically open populations being 

maintained by immigration of new breeders from outside the system. Yet, if researchers 

had gone no further than documenting the high densities of breeding birds at the site they 

would have erroneously concluded that no adverse effects on birds were occurring at 

Kesterson when in fact this reservoir was a highly dangerous “attractive nuisance” and 

regionally attracted waterbirds from other “source” populations into a demographic 

“sink” population. The trout populations in the Crow Creek watershed, just like the bird 

populations at Kesterson, are open populations with no barriers for movement of 

individuals into and out of the population. Accordingly, all of the investigative effort 

spent and space used in the Report to document trout densities and draw toxic risk 

conclusions from those densities [summarized on pp. 46-47] is misleading. These data 
simply show that trout are in the stream just as simple counts of breeding birds at 

Kesterson showed birds there were abundant. In neither instance would these data 

address the issue of whether the water quality was impaired, or not, from selenium or any 

other chemical. This comment also applies to the portion of the literature review that was 

devoted to reviewing the many other fish count studies that also erroneously conclude 

that high habitat occupancy, simply one measure of habitat attractiveness, equates to high 

habitat quality (i.e., low toxic risk). These studies fail to account for the mobility of the 

species and the confounding effect of immigration to data interpretation. Studies of that 

type, as opposed to the type represented by Vasterling (2003), are simply not designed to 

rule out the possibility that they are dealing with contaminant-induced demographic sink 

populations that would collapse in the absence of continual immigration from outside the 

site. Janz et al. (2010:191) recount a good case example of this issue. The measured 

whole body selenium concentrations in Thompson Creek trout were 4-14 ug/g dw and 

fish in this creek persisted across about 20 years of standardized monitoring. In a nearby 

tributary and adjacent pond the selenium concentration in trout averaged 13 ug/g dw 

whole body, yet the trout in these water bodies collapsed and disappeared within the first 

three years of monitoring. The primary difference between Thompson Creek and the 

nearby water bodies was Thompson Creek was open to fish movement into and out of the 

resident population from other areas, while the other nearby waters were not. 

5.0 LABORATORY STUDIES – EFFECTS CHARACTERISTICS

Because the proposed site-specific criterion is based on the brown trout toxicity study 

(Appendix F), review comments for this section will largely relate directly to the material 

presented in Appendix F of the Report and in Appendix A of Appendix F. 

There are two fundamental concerns with the interpretation of the brown trout toxicity 

study: 

The first concern is the lack of any true field control fish from which to calibrate point 

estimates of effects. The authors themselves state, “For the brown trout studies, the 

response observed at [hatchery] controls was adjusted to the response observed at 

background since true controls for this study were not practical.” [p. 50]

This is problematic given the evidence (discussed above) that the upstream Crow Creek 

and Deer Creek fish didn’t represent normal or background exposures to selenium, but 

probably something on the order of about 2-times to 4-times normal background 

exposures of 2-3 ug/g dw whole body fish tissue as found at the appropriate reference 

site. Until spawning fish can be collected from South Fork Tin Cup Creek or a 

comparably mining-independent reference site, with whole body tissue selenium < 3 ug/g 

dw and egg selenium < 4 ug/g dw (see Table 5-2 for South Fork Tin Cup Creek data), the 

point estimates of toxic effects cannot be confirmed to have been based on proper 

controls. If the fish sampled at upstream study area sites are compromised controls, then 

toxicity point estimates such as EC-10s or EC-20s will systematically overestimate the 

tissue concentration of selenium that causes a 10% or 20% toxic effect. To check for this 
potential source of systematic bias, comparison toxicity response curves could be 

calculated based on the hatchery controls where the data for hatchery fish seem suitable 

or on theoretical full performance values where the hatchery control data are problematic. 

The second concern is that the calculated point estimates of toxic effects are not based on 

the most sensitive reproductive endpoint. The point estimates are based on the endpoints 

of alevin survival to 15 days post-swim-up and sum total normal fry across four 

categories of deformity/abnormality. The EC-20s for those two endpoints are 21.6 ug/g 

and 21.7 ug/g egg selenium respectively (Appendix F: Table 8), yet Appendix F, 

Figures12 and 13, show that the EC-100 for swim-up occurs somewhere between about 

20 ug/g and 27 ug/g egg selenium. Although many of the brown trout alevins that had 

failed to swim-up by the time this study was terminated may have still been living (2.4-

33%), in nature total failure of swim-up would be equivalent to total reproductive failure. 

Analysis of the swim-up endpoint for less severe point estimates of effect than the EC100 should be explored. 

Related to the reported observation that “... yolk fry never swam up...” (Appendix F: 

page 17) for the five sets of eggs with the highest selenium exposures (also recorded as 0 

“% Swim-up” in Table 5-2 of the Report’s main body and illustrated that way in Figure 

12 of Appendix F), it is very perplexing as to how the 15-day Post Swim-up feeding trials 

produced any data for those same sets of eggs (reported in Appendix F, Figure 37)? 

Clearly, before a 15-day Post Swim-up Feeding Trial can be conducted alevins have to 

actually swim-up. If “... yolk fry never swam up...”, how is Post swim-up data 

presented in Figure 37 generated? 

The formulation of the “sum total normal fry” endpoint across four 

deformity/abnormality categories is not an environmentally realistic way to formulate 

that data [Appendix F, page 30 and Figure 44]. For example suppose five fry are each 

assessed for the four categories of deformity/abnormality (as per assessment methods, 

Appendix F, page 9) and the four effect categories are viewed once each, and each in a 

different one of four fry, leaving one fry that exhibits none of the deformity/abnormality 

effects. Then the “average” number of normal fry across the four categories is going to 

be 80%, 4 out of 5 fry will be free of any particular category of deformity/abnormality 

even though only one of the five fry in question, or 20%, will be a fully normal fry. 

That’s a large bias in the direction of underestimating toxic effect (and overestimating 

EC-10 and EC-20 concentrations) because environmental realism dictates that any fry not 

fully free of deformity/abnormality is unlikely to survive in nature. The deformity data 

ought to be re-analyzed based on a “proportion fully normal fry” endpoint. Such an 

endpoint would be expressed on a 0 to 1 scale rather than a 0 to 4 scale. 

When there is no causal relationship between two variables, as is the case with whole 

body selenium and egg selenium (dietary selenium is the causal variable for both), the 

proper procedure (e.g., see Gill 1987) to back translate from EC-10 and EC-20 egg 

selenium concentrations to associated whole body selenium concentrations in trout would 

be to directly calculate a new regression equation by reversing the original X and Y axes 

rather than by algebraically re-arranging the original regression equation to solve for 
whole body selenium as a function of egg selenium as the Report does on p. 65 of the 

main text and on p. 33 of Appendix F. Accordingly, the whole body translations from 

EC point estimates for egg selenium presented in the Report and in Appendix F warrant 

re-calculation. 

The species sensitivity distribution (SSD) illustrated in Figure 6-1 of the Report does not 

appear to be scaled properly on the Y-axis. For example at the brown trout EC-20 of 

21.6 ug/g dw egg selenium 1 of 7 species is affected or a proportion of 0.143, but the 

brown trout data point is plotted at about 0.080 on the Y-axis. Additionally, based on 

EPA’s current policy for tissue-based water quality criteria (i.e., Utah proposed sitespecific tissue-based criterion for the Great Salt Lake and 2010 draft proposed revision to 

the national chronic criterion) Figure 6-1 should be re-drafted based on EC-10 values 

rather than EC-20 values. If it is believed that results for all the other species of fish 

illustrated in Figure 6-1 are valid for comparison to brown trout results from this specific 

study area, then typically EPA would employ such data to estimate the 5th percentile 

value for the SSD as an acceptably protective chronic criterion for the entire 10-species 

fish community in the Smoky Canyon study area (of which only two species have been 

tested). If it is not believed all the results illustrated in Figure 6-1 would necessarily 

apply on a site-specific basis to the study area then the only site-specific sensitivity 

conclusion that can be reached is that brown trout are more sensitive than Yellowstone 

cutthroat trout. Thus, either a 5th percentile value should be calculated for the SSD 

presented in Figure 6-1, or a safety factor should be applied to the brown trout toxicity 

results to acknowledge that only two of ten site-specific species were tested for toxic 

sensitivity to selenium. 

6.0 UNADDRESSED ISSUES

 

Compliance with the Clean Water Act mandates water standards that will provide for the 

“... protection and propagation of fish, shellfish, and wildlife... (PL 92-500, Sec.101(2)). 

The Report is silent on all taxa other than fish and aquatic invertebrates. Recent research 

(Presser and Luoma 2010) reveals that fish whole body selenium concentrations are 

consistently about 1.1 times the selenium concentrations of co-located aquatic 

invertebrates. The Report asserts that the proposed site-specific fish egg criterion (21.6 

ug/g dw) would translate to about 14 ug/g fish whole body selenium, which, divided by 

1.1, would translate to about 12.7 ug/g dw aquatic invertebrate selenium. Based on 

Ohlendorf’s (2003: Figure 17.1) regression of mallard egg hatchability against dietary 

selenium exposure of hens the EC-10 is only 4.9 ug/g dw. Mallard hens, along with the 

majority of other species of water birds, rely heavily on aquatic invertebrates during the 

breeding season in order to meet the high protein requirements of ovulation. For a 

dietary exposure of 12.7 ug/g dw dietary selenium, Ohlendorf’s regression equation 

predicts a point toxicity estimate exceeding 85% failure of mallard eggs to hatch. Thus, 

it seems highly doubtful that the proposed site-specific criterion would comply with the 

Clean Water Act’s mandate to protect wildlife. As the studies of Skorupa et al. (2002) 

and Vasterling (2003) have documented, in both riparian and wetland systems in the 

phosphate-mining region of southeastern Idaho, avian eggs exhibit elevated selenium 
concentrations commensurate with the system wide levels of selenium contamination for 

species of birds across a spectrum ranging from blackbirds to mallards to geese to 

sandhill cranes. 

Employing brown trout as a sensitive surrogate for all the untested species of fish in 

Crow Creek, the goal would be to keep the selenium concentrations in eggs of all species 

below the brown trout-based site-specific criteria of 21.6 ug/g dw (for the moment 

neglecting whether this particular value is defensible or not, see recalculations below). 

This would be straight forward to monitor and manage if compliance monitoring were 

going to be based on direct measures of selenium in fish eggs. But for practical reasons 

(difficulty of sampling ripe female fish, as for example at the reference site) it is actually 

whole body selenium that is likely to be monitored. This raises the question of what 

whole body value would generally keep egg selenium below the target value in most (if 

not all) local species of fish? Determining this is not as simple as merely answering what 

whole body selenium concentration is likely to keep brown trout eggs below the target 

value and then extrapolating that to all other species of fish, as is done in the Report. The 

Report is silent on the rationale for this. Just as species of fish vary in their toxic 

sensitivity to selenium exposure, there is also a great deal of interspecies variability in the 

partitioning of selenium between different tissues. Studying a community of Colorado 

River Basin fish species, Osmundson and Skorupa (2011) found that simple ratios of egg 

to whole body selenium vary among species from average ratios of about 1 to 8. The 75th

percentile value was about 4. The Report found a ratio for brown trout, at the EC-20 

target for eggs of 1.54, i.e; the eggs contained about 1.54 times as much selenium as a 

whole body analysis would yield. Thus, a proposed egg target of 21.6 ug/g dw 

“translates” to a whole body target of (21.6)/(1.54) = 14 ug/g dw. 

However, a ratio of 1.54 is on the low end of the range of species variability for such 

ratios documented for 12 species of fish in the Colorado River Basin. If species 

variability of ratios is similar for the 10 species of fish in the Smoky Canyon study area, 

then a whole body target of 14 ug/g could result in most species of fish (60-70% in the 

Colorado River Basin data set) actually exceeding the 21.6 ug/g egg target (some species 

by a lot). The way to protect against this potential problem is to develop a site-specific 

set of tissue ratios for the 10 species of local fish (or most of them) and then select an 

appropriately protective translation ratio from that data set. Presumably such a local data 

set could be obtained within a spawning season or two of field sampling, as was achieved 

for the study of the Colorado River Basin fish community reported in Osmundson and 

Skorupa (2011). Alternatively, data from studies such as Osmundson and Skorupa 

(2011) could be employed to estimate a conservative ratio, for example Osmundson and 

Skorupa’s 75th percentile value. In any case, arbitrarily extrapolating the brown trout 

tissue translation ratio to all other species of fish is inappropriate and could lead to 

substantive loss of the presumed protectiveness of a site-specific criterion. 
SECTION II: Recommended Recalculations 

(1.) Hatch to15-days post swim-up fry survival endpoint: 

This toxicity response endpoint for the brown trout toxicity study (Appendix F of the 

Final Report) is based on raw data presented in Appendix A of the brown trout toxicity 

study (i.e., Appendix A of Appendix F) in Tables 3-3 and 3-4. The specific endpoint is 

the “Total Survival %” from Table 3-4 divided by the “% Hatch” from Table 3-3. For 

example, for control sample SPC-006 these figures would be [(96.0) / (99.7)] for a 

survival proportion of 0.963 from hatch until test termination at 15-days post swim-up. 

For reasons that this reviewer was unable to determine, 27 of the 36 reported survival 

proportions for this endpoint reported in Table 3-4 are incorrect. Of the 27 incorrect 

values, 24 are incorrect high and 3 are incorrect low. In some cases the discrepancy is 

quite substantial. For example, for sample LSV2C-004 the reported endpoint survival 

proportion is 0.662 (66.2%) even though the reported “Total Survival %” in Table 3-4 of 

16.8 and reported “% Hatch” from Table 3-3 of 50.6 actually yield a “hatch to 

termination” endpoint survival proportion of 0.332 (33.2%; i.e., 16.8 / 50.6). Thus, to 

begin with for this recommended recalculation the 27 incorrectly reported results were 

corrected. 

Next, there were five samples (LSV2C-003, LSV2C-004, LSV2C-005, LSV2C-010, and 

LSV2C-21) that did not produce any swim-up fry by the time the experiment was 

terminated. Instead of recording a post swim-up fry survival of 0.0% for those samples, 

the authors of the final report chose to record the percent survival among the 

developmentally “dead-ended” alevins. By doing this, they are assessing a post swim-up 

fry endpoint for most of the samples and a developmentally completely different deadend alevin endpoint for the five samples mentioned above. Mixing two different 

endpoints for exposure-response modeling fundamentally invalidates the modeling. This 

error is easily corrected by using only the post swim-up fry endpoint and recording a post 

swim-up fry survival of 0.0% for the five samples in question. 

The authors excluded data from the hatchery control samples because the first set of 

hatchery controls performed very poorly with regard to % hatch and the second set of 

hatchery controls were obtained as eyed embryos and therefore weren’t comparable to the 

starting point for all the other samples (started as undeveloped fertilized eggs). While the 

second set of controls yielded excellent % hatch performance, that may have been due to 

the advanced stage of the embryos at receipt and thus much shorter holding time to hatch 

(9-13 days) than for any of the other samples (which had holding times to hatch of 37-47 

days). In addition to those data exclusions, there were 10 additional samples with 

missing test organisms because of clogged drain pipes that caused water to overflow out 

of the test system, apparently carrying some test organisms away (see Appendix A of 

Appendix F). The authors do not know how many of the carried away organisms were 

dead or alive, deformed or normal, or anything else about them. The authors chose to 

simply ignore the missing test organisms, effectively assuming the proportion of missing 
fish that were live or dead and deformed or normal was exactly the same as those 

proportions in the fish left behind. There is no valid justification for such an assumption. 

Best scientific protocol dictates that the uncertain results for these 10 samples also be 

excluded from exposure-response modeling. 

Thus, the valid data set for modeling the “hatch to15-days post swim-up fry survival” 

endpoint is as follows (Table 1; log transformation of exposure data is performed on ppb 

exposure instead of ppm exposure to avoid negative logarithms): 

Table 1. Corrected dataset for the hatch to 15-days post swim-up fry survival endpoint. 

[1] 

Sample ID 

[2] 

Egg Se 

ppm dry 

wgt. 

[3] 

 Log Egg 

Se ppb dry 

wgt. 

[4] 

Total (post 

swim-up 

fry) 

Survival 

% 

[5] 

 % Hatch 

[6] 

Hatch to 

15-day 

post swimup 

proportion 

survival 

[7] 

Number 

hatched 

[8] 

Expected 

number of 

surviving 

fry 15-

days post 

swim-up 

[6] X [7] 

CC-150-

009 

12.8 4.107 27.0 28.5 .947 171 162 

CC-150-

011 

8.4 3.924 95.3 96.0 .993 288 286 

CC-150-

012 

8.5 3.929 86.0 88.8 .967 311 301 

CC-150-

015 

9.1 3.959 77.5 78.3 .990 470 465 

CC-150-

018 

6.9 3.839 84.3 87.2 .967 523 506 

CC-150-

020 

6.2 3.792 96.3 97.3 .990 584 578 

CC-350-

006 

14 4.146 67.7 71.7 .944 430 406 

LSV2C003 40.3 4.605 0.0 93.5 .000 374 0 

LSV2C004 36 4.556 0.0 50.6 .000 253 0 

LSV2C005 26.8 4.428 0.0 71.3 .000 214 0 

LSV2C010 38.8 4.589 0.0 87.0 .000 87 0 

LSV2C012 13.2 4.121 95.7 98.3 .974 590 575 

LSV2C016 13.4 4.127 91.7 95.0 .965 570 550 

LSV2C021 28.1 4.449 0.0 69.3 .000 416 0 
The authors of the Report employ EPA’s Toxicity Relationship Analysis Program 

(TRAP) for modeling exposure-response. However, TRAP offers several options with 

regard to statistical models. The authors chose to use the Least-Squares Nonlinear 

Logistic Regression model (see Appendix E of Appendix F). EPA advises in the TRAP 

program overview that the Least Squares Logistic Regression model is most appropriate 

for endpoints that are (1) measured on a continuous scale (for example weight or height) 

and (2) expected to have infinite tails approaching 0% and 100% response. Neither of 

these conditions applies to the brown trout data in question. The data in question are (1) 

non-continuous dichotomous (binomial) data (each hatched egg either goes on to become 

a 15-day post swim-up fry or does not) and (2) the endpoint would be expected to have a 

finite “threshold” response in which there is no response until a minimum finite threshold 

exposure is experienced, not an infinitely gradual response tail. For dichotomous, 

threshold response data, EPA advises that the most appropriate TRAP model choice is the 

Tolerance Distribution Analysis fitted to a Triangular Distribution. Therefore, for this 

recalculation those are the modeling options that were chosen for TRAP exposureresponse modeling with the following result (Figure 1): 

Figure 1. Trap output for corrected hatch to 15-day post swim-up fry survival endpoint. 
The lack of useable data points within the intermediate response range violates the TRAP 

program requirements for model fitting, generating an error warning message. The EC10 and EC-20 estimates of 14.9 and 16 ppm shown above are the result of forcing a 

model fit and should not be viewed as reliable. Recalculation for this endpoint reveals 

that the useable dataset is not sufficient for reliably estimating any EC-10 or EC-20 

values. The only reasonably certain information that these data yield is that the finite 

threshold for total response, the EC-100 threshold, occurs somewhere between an egg 

selenium concentration of 14 ppm and 27 ppm. Because 100% response (i.e., total failure 

of brown trout alevins to achieve swim-up) is already being expressed at an egg selenium 

concentration of 27 ppm, and because valid estimates of the EC-10 and EC-20 cannot be 

modeled from the useable dataset, proposing an egg selenium standard of nearly 22 ppm 

(as the authors do in the final report) based on an invalid EC-20 estimation (as the authors 

do in the Report) does not seem very defensible. Any site-specific standard proposal 

higher than 14 ppm egg selenium cannot be defensibly supported by the “hatch to 15-day 

post swim-up fry survival” endpoint. 

(2.) Larval Deformity Endpoint: 

The collection of larval deformity data was seriously flawed in at least two ways. 

First, the larvae assessed for deformities were not a random subsample of all larvae. 

Alevins and post swim-up fry that died during the brown trout toxicity study mostly did 

not get assessed for presence or absence of deformities (see Section 2.5 of Appendix A to 

Appendix F). Since deformed larval fish can be expected to have a differentially high 

propensity to die compared to undeformed larval fish, all the measures of percent 

deformed larvae are likely biased low. For example, among a limited number of test 

groups (samples LSV2C-004, LSV2C-005, and LSV2C-021) deformity assessments are 

available for reasonable samples of both dead and live larval fish as follows (Table 2 

from raw data presented in the first three of the un-numbered Tables in Appendix C of 

Appendix F): 

Table 2. Comparison of percent deformed fry in samples of dead and live specimens 

(sample sizes in parentheses). 

[1] 

Sample ID 

[2] 

 Type of deformity 

[3] 

 % deformed dead larval 

fish (n) 

[4] 

% deformed live larval 

fish (n) 

LSV2C-004 Craniofacial 100 (101) 1.6 (64) 

LSV2C-004 Vertebral 100 ( 64) 68.3 (63) 

LSV2C-005 Craniofacial 100 ( 89) 44.9 (49) 

LSV2C-005 Vertebral 100 ( 84) 65.3 (49) 

LSV2C-021 Craniofacial 100 (101) 31.9 (69) 

LSV2C-021 Vertebral 100 ( 62) 71.0 (69) 
As the above comparisons show, the bias introduced by assessing mostly live (surviving) 

larval fish for deformities can be substantive. However, the comparisons shown above 

are for test samples that had high selenium exposures (27-36 ppm egg selenium) and 

presumably therefore would express this type of bias to a greater extent than test samples 

from lower exposures would. Presumably, lower exposures would not uniformly produce 

exclusively (100%) deformed dead larval fish. Since the authors did assess greater 

numbers of dead fish for the five highest exposure groups and the lower exposure groups 

often yielded predominantly surviving larval fish, it is possible that the overall magnitude 

of this bias on exposure-response modeling for larval deformities is relatively modest. 

Nonetheless, the main points to consider are that the exact amount of bias is unknown, 

and whatever the magnitude of bias is it is certain to be in only one direction, 

underestimation of response values (thus overestimation of toxic threshold 

concentrations). 

Second, each of the four types of deformity examined were assessed and reported 

completely independently of each other (see Appendix C of Appendix F). Thus, while it 

is possible to determine what percent of assessed larval fish were normal for each type of 

deformity, it is not possible to determine what percent of assessed larval fish were 

completely undeformed (i.e., were fully “normal” larval fish). Unfortunately, the raw 

deformity assessment data were not available for review, but were cited (see p. 3-10 of 

Appendix A of Appendix F) as being presented in a separate document that Simplot 

Company refused to release for this review. The authors of the Report used an endpoint 

that they called “sum total normal fry” which in reality should more accurately have been 

called “sum partial normal fry” because it is unknown if any of the fry in this sum are 

simultaneously normal for all four types of deformity. It is only known that the fry in this 

sum are normal for at least one of the four types of deformity (craniofacial, vertebral, fin, 

edema). Of course the sum of partial normal fry will necessarily overestimate the sum of 

fully normal fry and it is only fully normal fry that can be expected in nature to be free of 

any deformity-related toxic impairment. A recalculation that can be done here to 

partially correct for the overestimation of fully normal fry is to recognize that the 

proportion of fully normal fish cannot have been any higher than the lowest proportion of 

normal fish in any of the four deformity categories reported in Appendix C of Appendix 

F. This will be referred to as “maximum fully normal fry”. That endpoint should be a 

much better approximation of percent fully normal fry than the author’s “sum partial 

normal fry” endpoint. Also, for this recalculation the second set of hatchery controls 

provide useable estimates of expected background deformity rates because the hatching 

rates were all acceptable control values (i.e., >90% hatch) and from hatch to 15-days post 

swim-up fry these controls are strictly comparable in their holding times to the fieldexposed test groups. However, the results from the “accidental overflow” test samples 

must again be excluded from analysis. This yields the following data set (Table 3) for 

modeling the response variable of larval deformity (log transformation of exposure data 

is performed on ppb exposure instead of ppm exposure to avoid negative logarithms): 
Table 3. Dataset for maximum normal fry endpoint. 

[1] 

 Sample ID 

[2] 

Egg Se ppm 

dry wgt. 

[3] 

 Log Egg Se 

ppb dry wgt. 

[4] 

Maximum 

proportion 

totally 

undeformed 

(normal) fry 

[5] 

Number 

undeformed fry 

for most 

common of 

four types of 

deformity 

[6] 

Number of 

assessed fry for 

most common 

of four types of 

deformity 

CC-150-009 12.8 4.107 .768 109 142 

CC-150-011 8.4 3.924 .429 114 266 

CC-150-012 8.5 3.929 .637 191 282 

CC-150-015 9.1 3.959 .519 231 445 

CC-150-018 6.9 3.839 .593 288 486 

CC-150-020 6.2 3.792 .894 499 558 

CC-350-006 14 4.146 .513 198 386 

LSV2C-003 40.3 4.605 .000 0 98 

LSV2C-004 36 4.556 .158 20 127 

LSV2C-005 26.8 4.428 .128 17 133 

LSV2C-010 38.8 4.589 .000 0 71 

LSV2C-012 13.2 4.121 .424 235 554 

LSV2C-016 13.4 4.127 .915 485 530 

LSV2C-021 28.1 4.449 .153 20 131 

SPC-001 0.73 2.863 .863 490 568 

SPC-003 0.73 2.863 .822 448 545 

SPC-005 0.73 2.863 .848 476 561 

SPC-006 0.73 2.863 .854 475 556 

The authors of the Report employ EPA’s Toxicity Relationship Analysis Program 

(TRAP) for modeling exposure-response. However, TRAP offers several options with 

regard to statistical models. The authors chose to use the Least-Squares Nonlinear 

Logistic Regression model (see Appendix E of Appendix F). EPA advises in the TRAP 

program overview that the Least Squares Logistic Regression model is most appropriate 

for endpoints that are (1) measured on a continuous scale (for example weight or height) 

and (2) expected to have infinite tails approaching 0% and 100% response. Neither of 

these conditions applies to the brown trout data in question. The data in question are (1) 

non-continuous dichotomous (binomial) data (each hatched egg either goes on to become 

an undeformed fry or a deformed fry) and (2) the endpoint would be expected to have a 

finite “threshold” response in which there is no response until a minimum finite threshold 

exposure is experienced, not an infinitely gradual response tail. For dichotomous, 

threshold response data, EPA advises that the most appropriate TRAP model choice is the 

Tolerance Distribution Analysis fitted to a Triangular Distribution. Therefore, for this 

recalculation those are the modeling options that were chosen for TRAP exposureresponse modeling with the following result (Figure 2): 
 Figure 2. Trap output for maximum fully normal fry endpoint. 

This dataset met all TRAP program requirements for a clean model fit without any error 

warnings. The estimated EC-10 is 6.5 ppm egg selenium and the estimated EC-20 is 9.0 

ppm egg selenium. These estimates are not fully corrected for the “sum partial normal 

fry” bias compared to proportion fully normal fry and therefore are still overestimates of 
the EC-10 and EC-20 for the larval deformity endpoint. Even so, both of these 

recalculated estimates are substantially lower than the site-specific standard of almost 22 

ppm egg selenium proposed by the authors of the Report. This endpoint also illustrates 

why it is so crucial to eventually obtain valid field control data. If the hatchery control 

data points are omitted, and it is erroneously assumed that the CC-150 and CC-350 

(upstream) sampling sites are valid field control sites then the estimates of the EC-10 and 

EC-20 for larval deformity respectively increase to 15.2 and 17.5 ppm egg selenium as 

illustrated below (Figure 3): 

 Figure 3. TRAP output for maximum fully normal fry endpoint excluding 

 the hatchery control data points. 
Obtaining valid field control results for the larval deformity endpoint should be a very 

high priority for future research. 

(3.) Regressing X on Y for Estimation of Critical Whole Body Tissue Concentrations: 

Because of the distinctly species-specific nature of selenium partitioning between tissue 

types within individual fish (Osmundson and Skorupa 2011) and because of the complex 

spatio-temporal biology of selenium deposition into brown trout eggs which can be 

expected to differ in a hatchery versus in the field (Estay et al. 2003), this recalculation is 

based only on the field data for brown trout as presented in Table 5-2 of the final report. 

Because tissue concentrations from polluted environments typically conform to lognormal distributions, the regression recalculated here is calculated on log-log transformed 

data. The recalculated regression is illustrated below (Figure 4). The regression is highly 

significant (p < 0.0000001) and the adjusted R-square value for the regression is 0.788. 

Log-Log Regression of Brown Trout WB Selenium (dry wgt.) on Egg Selenium 

(dry wgt)

Log WB Se = 0.199 + 0.668 x Log Egg Se

0.7 0.8 0.9 1.0 1.1 1.2 1.3 1.4 1.5 1.6 1.7

Log Egg Se

0.6

0.7

0.8

0.9

1.0

1.1

1.2

1.3

1.4

Log WB Se

Figure 4. Regression of X on Y for back-calculating critical whole body concentrations 

from EC-10 and EC-20 egg concentrations. 
This regression yields a critical whole body selenium concentration of 12.32 ppm at an 

egg selenium concentration of 21.63 ppm, a result that is slightly lower than the value of 

13.35 ppm critical whole body selenium reported for brown trout data in the Report based 

on algebraic re-arrangement of the egg selenium on whole body selenium regression 

equation (see p. 65). When the recalculated regression above (Figure 4) is evaluated at 

the EC-10 for maximum fully normal fry (Figure 2) of 6.5 ppm egg selenium it yields an 

estimated critical whole body selenium concentration of 5.5 ppm. 

(4.) Test of Assumption that the 14 ppm Proposed Critical Whole Body Tissue 

Concentration Derived from Brown Trout Tissue Partitioning Data Would 

Generally Keep Egg Selenium Values Below a Critical Value of 21.63 ppm 

Across a Multi-species Community of Fish: 

Tissue partitioning ratios from elevated selenium exposure can be expected to be dosedependent. Therefore to examine this issue using the fish community dataset from the 

Colorado River Basin reported in Appendices 1 and 2 in Osmundson and Skorupa (2011), 

only the fish samples with egg selenium results relatively near (in the range of 15 to 30 

ppm) the target value of 21.63 ppm proposed in the Report were evaluated. The results 

are presented below (Table 4): 

Table 4. Expected egg selenium concentrations at a whole body selenium concentration 

of 14 ppm in a multi-species fish community (exceedances of proposed criterion in bold 

type). 

Fish Species Ovary/Egg Se Whole Body Se OE:WB Ratio Expected Egg Se 

at Whole Body Se 

of 14 ppm 

Green Sunfish 27.4 22.8 1.20 16.8 

Green Sunfish 21.8 15.8 1.38 19.3 

Green Sunfish 18.1 11.9 1.52 21.3 

Green Sunfish 15.2 9.1 1.67 23.4

Green Sunfish 15.2 9.7 1.57 22.0 

Common Carp 16.3 11.7 1.39 19.5 

Common Carp 27.3 23.1 1.29 18.1 

Roundtail Chub 15.2 6.44 2.36 33.0 

Roundtail Chub 17.8 8.4 2.12 29.7 

Roundtail Chub 18 6.6 2.73 38.2 

Roundtail Chub 16.9 7.2 2.35 32.9 

Black Bullhead 26.4 8.59 3.07 43.0 

Black Bullhead 28.6 2.9 9.86 138 

Channel Catfish 30.3 4.0 7.58 106 

Channel Catfish 21.1 3.3 6.39 89.5 

Channel Catfish 29.5 3.4 8.68 122 

Channel Catfish 15.9 1.9 8.37 117 

Channel Catfish 15.2 2.4 6.33 88.6 
Thus, for the Colorado River Basin community of fish, assuming that a critical whole 

body concentration of selenium derived from brown trout data would be universally 

applicable would have allowed other species of fish to reach egg selenium concentrations 

as high as 138 ppm, or more than 6-times the proposed site-specific standard of 21.63 

ppm. Moreover, the overall exceedance rate would be 68% (weighting equally each 

species exceedance rate in Table 4). In other words, for this example fish community, a 

whole body tissue criterion of 14 ppm (that would keep brown trout eggs under 21.63 

ppm) would keep only 32% of other fish species eggs under 21.63 ppm. Consequently, 

in this case, assuming that all fish species partition selenium between tissues comparably 

to brown trout would be a very under protective assumption. This exercise does not 

prove that such an assumption would also be very under protective as applied to the Crow 

Creek fish community, but it does emphasize that making such an assumption for the 

Crow Creek fish community (as is done in the Report) would seem unlikely to hold true 

and therefore must be validated before it can be accepted for regulatory purposes. 

REVIEWER’S SUMMARY

1. Comparative data for trout versus comparative data for water and periphyton from 

the reference site (outside the mining district) and the purported background 

upstream study area sites suggest that regionally appropriate background whole 

body tissue concentrations should be about 2-3 ug/g dw rather than the 6-9 ug/g 

dw measured at the purported background upstream sites. 

 

2. The population level surveys of fish abundance throughout the study area do not 

account for the ability of fish to freely move into the studied populations from 

outside the study area. 

3. Because the upstream study area sampling sites did not represent normal 

background conditions with respect to trout, and no spawning trout from the 

reference site were available for sampling, the laboratory toxicity tests for brown 

trout were conducted without a valid field control. 

4. It is uncertain that the most sensitive endpoint was evaluated for threshold toxicity 

point estimates. Complete failure (EC-100) of alevin swim-up appears to have 

been reported at egg selenium concentrations comparable to the EC-20 for the 

Report’s chosen endpoints. 

5. One of the chosen endpoints, sum total normal fry, is formulated in an 

environmentally unrealistic manner that is essentially the sum of partially normal 

fry. 

6. The Species Sensitivity Distribution in the Report should be re-formatted based 

on EC-10s rather than EC-20s to be consistent with current USEPA policy 

regarding tissue-based criteria, and analyzed for its 5th percentile value. 
7. How the proposed site-specific criterion would meet the Clean Water Act 

mandate for protection and propagation of wildlife should be addressed. 

8. It is recommended that translation of a fish egg selenium target to a fish whole 

body selenium target not be arbitrarily extrapolated from brown trout to all other 

species of fish without estimating how protective that would be for other species 

of fish. 

9. Correction of errors in the “hatch to 15-day post swim-up fry survival” endpoint 

database and re-analysis of the corrected data reveal that the data are insufficient 

for reliably estimating EC-10 and EC-20 tissue concentrations. What can be 

concluded with high confidence from these data is that the EC-100 is < 27 ppm 

egg selenium (dw). 

10. The “larval deformity” endpoint appears to be a more sensitive toxicity endpoint 

than the “hatch to 15-day post swim-up fry survival” endpoint. However, data 

collection for this endpoint suffers from lack of random sampling, and clear 

presentation of the percent fully undeformed (normal) fry. To the extent that 

these shortcomings could be partially corrected for, a re-analysis of the deformity 

data yielded estimated EC-10 and EC-20 values < 10 ppm egg selenium. This is 

substantively lower than the draft proposed site-specific criterion of 21.63 ppm. 

11. The proposed critical whole body selenium tissue concentration of 14 ppm, based 

on tissue partitioning dynamics in brown trout, was found to be only 32% 

protective when applied to a full community of Colorado River fish species. It 

therefore seems unlikely that assuming all species of fish partition selenium 

between different tissues comparably to brown trout in the Crow Creek fish 

community will hold true. 

 

12. At the re-calculated EC-10 concentration for larval deformity of 6.5 ppm egg 

selenium, the estimated critical whole body selenium concentration is 5.5 ppm, an 

outcome substantially lower than the draft proposed site-specific criterion of 14 

ppm. 
Literature Cited

Besser, J., C. Ivey, J. Kunz, B. Brumbaugh, and C. Ingersoll. 2008. Toxicity of dietary 

and waterborne selenium to desert pupfish in a life-cycle toxicity test. U.S. 

Geological Survey, Columbia, MO. Presented at the 29th Annual North America 

Meeting of the Society of Environmental Toxicology and Chemistry (SETAC). 

Tampa, FL (November 16-20). 

Birkner, J.H. 1978. Selenium in aquatic organisms from seleniferous habitats. Ph.D. 

Thesis. Colorado State University, Fort Collins, CO. 

Canton, S., and D. DeForest. 2009. Technical Memorandum to the National Mining 

Association and Rio Tinto – Review of Selenium Concentrations in Desert 

Pupfish and Comments on Fecundity as a Selenium Toxicity Endpoint for Fish. 

Estay, F., A.Diaz, R. Pedrazza, and N. Colihueque. 2003. Oogenesis and plasma levels 

of sex steroids in cultured females of brown trout (Salmo trutta linneaus 1758) in 

Chile. J. Exp. Zool. Part A: Comp. Exp. Biol., 298A:60-66. 

Gill, J.L. 1987. Biases in regression when prediction is inverse to causation. J. Anim. 

Sci., 64:594-600. 

Hamilton, S.J., and V.P. Palace. 2001. Assessment of selenium effects in lotic 

ecosystems. Ecotoxicol. Environ. Safety, 50:161-166. 

Hansen, L.D., K.J. Maier, and A.W. Knight. 1993. The effect of sulfate on the 

bioconcentration of selenate by Chironomus decorus and Daphnia magna. Arch. 

Environ. Contam. Toxicol., 25:72-78. 

Janz, D.M., D.K. DeForest, M.L. Brooks, P.M. Chapman, G. Gilron, D. Hoff, W.A. 

Hopkins, D.O. McIntyre, C.A. Mebane, V.P. Palace, J.P. Skorupa, and M. 

Wayland. 2010. Selenium toxicity to aquatic organisms. Pp. 141-231 In: P.M. 

Chapman, W.J. Adams, M.L. Brooks, C.G. Delos, S.N. Luoma, W.A. Maher, 

H.M. Ohlendorf, T.S. Presser, and D.P. Shaw (eds.), Ecological Assessment of 

Selenium in the Aquatic Environment. CRC Press, Boca Raton, FL. 

Kennedy, C.J., L.E. McDonald, R. Loveridge, and M.M. Strosher. 2000. The effects of 

bioaccumulated selenium on mortalities and deformities in the eggs, larvae, and 

fry of a wild population of cutthroat trout (Oncorhynchus clarki lewisi). Arch. 

Environ. Contam. Toxicol., 39:46-52/ 

Lemly, A.D. 1993. Metabolic stress during winter increases the toxicity of selenium to 

fish. Aquat. Toxicol., 27:133-158. 
Lowe, T.P., T.W. May, W.G. Brumbaugh, and D.A. Kane. 1985. National Contaminant 

Biomonitoring Program: concentrations of seven elements in freshwater fish, 

1978-1981. Arch. Environ. Contam. Toxicol., 14:363-388. 

Ogle, R.S., and A.W. Knight. 1996. Selenium bioaccumulation in aquatic ecosystems: 1. 

Effects of sulfate on the uptake and toxicity of selenate in Daphnia magna. Arch. 

Environ. Contam. Toxicol., 30:274-279. 

Ohlendorf, H.M. 2003. Ecotoxicology of selenium. Pp. 465-500 In: D.J. Hoffman, B.A. 

Rattner, G.A. Burton, Jr., and J. Cairns, Jr., (eds.), Handbook of Ecotoxicology 

(2nd Ed.). CRC Press, Boca Raton, FL. 

Osmundson, B., and J. Skorupa. In review. Selenium in fish tissue: prediction equations 

for conversion between whole body, muscle, and eggs. Final Rpt., Environmental 

Contaminants Program, Off-Refuge Investigations. U.S. Fish and Wildlife 

Service, Grand Junction, CO. 

Presser, T.S., and S.N. Luoma. 2010. A methodology for ecosystem-scale modeling of 

selenium. Integr. Environ. Assess. Manage., 6:685-710. 

Skorupa, J.P. 1998. Selenium poisoning of fish and wildlife in nature: lessons from 

twelve real-world experiences. Pp. 315-354 In: W.T. Frankenberger, Jr., and R.A. 

Engberg (eds.), Environmental Chemistry of Selenium. Marcel Dekker, New 

York, NY. 

Skorupa, J.P., S.J. Detwiler, and R. Brassfield. 2002. Reconnaissance survey of 

selenium in water and avian eggs at selected sites within the phosphate mining 

region near Soda Springs, Idaho, May-June 1999. Final Rpt., Environmental 

Contaminants Program, Off-Refuge Investigations. U.S. Fish and Wildlife 

Service. Sacramento, CA 

Vasterling, R.P. 2003. Selenium contamination effects on avian reproductive success 

and population dynamics in phosphate-mining regions of southeastern Idaho. 

Masters Thesis, University of Idaho. 

Walsh, D.F., B.L. Berger, and J.R. Bean. 1977. Residues in fish, wildlife, and estuaries: 

mercury, arsenic, lead, cadmium, and selenium residues in fish, 1971-73 --- 

National Pesticide Monitoring Program. Pest. Monit. J., 11:5-34. 

Williams, M.J., R.S. Ogle, A.W. Knight, and R.G. Burau. 1994. Effects of sulfate on 

selenate uptake and toxicity in the green alga Selenastrum capricornutum. Arch. 

Environ. Contam. Toxicol., 27:449-453. 
APPENDIX 1. 

Peer Reviews and Peer Reviewer Curriculum Vitae 
Peer Review Scorecard 

Scientific Issue Peer 

1 

Peer 

2 

Peer 

3 

Premature to interpret uncompleted YCT toxicity testing study + + (-) 

Current EPA chronic water quality criterion for Se not based solely on 

aqueous exposure 

+ + + 

EPA’s draft 2004 chronic criterion not based on Belew’s Lake data + ++ + 

Follow-up EPA sponsored winter stress study did not replicate Lemly + (-) ++ 

The literature review is distinctly uneven + ++ 0 

Sulfate inhibition of selenium bioaccumulation a nonfactor 0 + + 

Invalid assumption that upstream sample sites represent background (+) + ++ 

Brown trout vitellogenesis chronology not properly considered + ++ 0 

Fish densities and age class distributions invalid indicators of effects (+) ++ ++ 

Absence of true field controls for toxicity testing (+) + + 

Swim-up failure EC-100 not properly noted + + + 

Frequency of fully normal fry is the proper deformity endpoint + + + 

X on Y regression preferable to algebraic rearrangement of Yon X for 

back calculating from egg Se to whole-body Se + 0 0 

SSD curve should be based on EC-10s not EC-20s 0 (+) 0 

Wildlife risk assessment required, but not conducted + + 0 

Inappropriate extrapolation of Brown Trout tissue partitioning ratio ++ (-) ++ 

Correction required for 27 hatch to post swim-up survival data points ?? + ?? 

Swim-up failure beakers should have been counted as 0% survival + + + 

Replicates w/ missing organisms should be excluded from analyses + + ++ 

Inappropriate TRAP option selected for modeling + (+) (+)

15-day post swim-up survival endpoint results unsuitable for modeling (+) + 0 

Survivor bias in larval deformity assessments (+) + + 

“Maximum fully normal fry” adjustment warranted + + 0 

EC-10 & EC-20 estimates substantively lower than reported + ?? 0 

Key: ++ = concurrence & expand point beyond FWS critique; + = concurrence; (+) = 

conditional concurrence; 0 = no judgment; (-) = conditional non concurrence; - = non 

concurrence ; ?? = peer reviewer too confused by data in draft report to make judgment 
Page 1

I (Mark Rigby, Ph.D.) was asked by Atkins Global to review USFWS comments on the “Interpretive

Findings for Field and Laboratory Studies and Literature Review in Support of a Site‐Specific Selenium

Criterion, Smoky Canyon Mine.”  The Interpretive Report represents a substantial body of research and

effort.  USFWS has reviewed the report and provide thoughtful comments with which I generally concur.   

I have been performing human health and ecological risk assessments in the United States since 2000.  

Those risk assessments have mainly been on behalf of the military bases.  With respect to selenium, I

have performed risk assessments for fishes and aquatic birds on behalf of the Salton Sea Authority for

solar evaporation ponds and other surface water bodies in the Salton Sea Area.  I also worked on the

North San Francisco selenium TMDL while I was employed at Tetra Tech.  Recently, I have been working

with Dennis Lemly of USFWS on re‐evaluating several published papers on the toxicity of selenium to

various fishes.  

My evaluation of USFWS’ comments follows.

Responses to Comments from USFWS.

1. Introduction

A. Interpretive findings for Yellowstone cutthroat trout (YCT) are brought into question by the

disclaimer presented in this section stating that this “... Interpretive Report precedes the

completion of the YCT studies reporting...” If the YCT toxicity studies are not yet completed,

then, by normal scientific protocol, it would be premature to attempt to interpret those studies.

[p. 3]

Evaluation:  Concur.  The interpretive report should be based on the final results of the YCT

toxicity testing.   Although Appendix H is labeled as “draft,” Appendix G appears to be final.

B. The Report exhibits some confusion regarding the basis for EPA’s current national water quality

chronic criterion for selenium of 5 parts per billion (ug/L). It states that it is “...based on aqueous

exposure only and includes no dietary component of exposure...” which is erroneous. The Report

authors correctly state that the criterion is based on field data from Belews Lake in North

Carolina, but they fail to link that with the fact that fish at Belews Lake experienced both

aqueous and dietary exposure to selenium as is typical of a field study at any selenium‐polluted

lake. [p. 5]

Evaluation:  Concur.

C. The Report again exhibits a degree of confusion with regard to the basis for EPA’s draft chronic

criterion revision of 2004 stating that it also was based on data from Belew’s Lake in North

Carolina. It was not; it was based on a laboratory experiment (Lemly 1993). To add to this

confusion, the authors then ask the reader to compare GMCVs (genus mean chronic values) for

trout from EPA’s 2004 draft document to the lower bluegill‐based draft criterion proposal while

failing to mention that the bluegill result is for an entirely different, and more sensitive, toxicity

test (multiple stressors of Se and temperature) than the trout GMCVs (single stressor, Se only).

Thus, the implication that trout are more tolerant than bluegill is based on a false comparison.

[p. 6]

Evaluation:  Concur.  As written, the text is incorrect.  The Interpretive Report should

provide a table comparing chronic toxicity values for trout and bluegill from the literature to
Page 2

illustrate their point.  However, at first glance, the studies reviewed by USEPA (2010) do not

appear to support this point.   

D. It is correctly noted in the Report that USEPA commissioned a follow‐up study meant to

replicate and extend the work of Lemly (1993), but the Report fails to note that the follow‐up

study did not replicate Lemly’s work. Lemly’s work focused on the interactive effects of

selenium exposure and winter‐stress syndrome (a physiological response to declining water

temperature and decreasing photo period expressed as lipid loss). The follow‐up study did not

decrease photo period and only lowered water temperature, which failed to induce lipid loss

(i.e., failed to induce winter‐stress syndrome). Because the follow‐up study did not mimic

“winter”, unlike Lemly’s work, it does not refute or replace Lemly’s original work; furthermore

this limited study is of unknown applicability to free‐ranging fish. [p. 6]

Evaluation:  Concur.  While I do not disagree, it should be noted photoperiod and the

applicability of USEPA (2008) to field populations was not discussed by USEPA (2008).  

Therefore, this comment is based on USFWS’ review and interpretation of the USEPA (2008)

study.  Nonetheless, USFWS is correct when they state that USEPA (2008) did not

manipulate photoperiod and, therefore, the results may not be applicable to fish outside of

the laboratory.  The Interpretive Report should be revised to include USFWS’ critique of

USEPA (2008).

2. Site setting

A. The field sampling design is predicated on an assumption that sampling locations upstream of

selenium discharges to Crow Creek will represent water quality conditions “...unimpacted by

mining activities.” However, as will be discussed below, the biological data contradict that

assumption, which has important implications for the laboratory trout toxicity studies (also

discussed below). [p. 11]

Evaluation:  see response to Comment 4.B.

3. Literature review

A. The literature review is distinctly uneven. For example, Besser et al.’s (2008) reporting of a low

toxicity threshold for desert pupfish is dismissed based on an unpublished memo (Canton and

DeForest 2009) prepared on behalf of the Mining Association and Rio Tinto (a mining

corporation); a memo that is not easily accessible to the scientific community, as no URL link to

a public posting of this privately held memo is provided in the References section of the report

[p. 22 and p. 74]. Yet, results that support a high toxicity threshold reported by Kennedy et al.

(2000) are accepted uncritically with no mention of a peer‐reviewed, published critique of

Kennedy et al. (2000) by Hamilton and Palace (2001) [p. 24 and p. 25]. Hamilton and Palace

argued, for example, that the study was compromised by unexplained high mortality in eggs

obtained from the reference site and by such a narrow exposure interval among most non‐

reference eggs that there would be “... little hope of deciphering [statistically] significant

differences in reproductive success.” Issues also could be taken with many other specifics of the

literature review, but ultimately the Report’s derivation of a proposed site‐specific selenium

criterion is not based substantively on the literature review, so the limited time available for this

review will be allocated to higher priority comments.

Evaluation:   
Page 3

i. The Interpretive Report states that its literature review is focused on “relevant studies

of cold water species”, which is entirely appropriate.  However, the Interpretive Report

should define what “relevant” species means.  I assume that this refers to those species

present in the study area.

ii. USFWS is correct that the review by Canton and DeForest (2009) is not available online

and is not in a peer‐reviewed journal.  However, if the literature review is focused on

species that occur in the study area, then the discussion of selenium toxicity to desert

pupfish should be removed, as they do not occur in the study area.  If Besser et al.

(2008) and Canton and DeForest (2009) are going to be discussed in the literature

review, the Interpretive Report should provide both in their entirety as an Appendix.  

Further, all unpublished studies (e.g., the reports by GEI Consultants) should be included

in their entirety in an Appendix.

iii. A number of the studies reviewed in the Interpretive Report are rather problematic in

experimental design, data analysis, and interpretation.  Therefore, without careful

review and re‐analysis of the original data, the findings of many of the studies included

in the literature review may be of limited utility.  A critique and evaluation of all of the

studies included in the literature review would be very lengthy and time consuming, as

indicated by USFWS.  However, USEPA (2010) has done some of that work and USEPA’s

(2010) review of the literature should probably be included in the Interpretive Report.  It

should also be noted that the study by Golder from 2009 on Dolly Varden char was

published in 2010 (McDonald et al. 2010.  Environmental Toxicology and Chemistry

29:2800–2805) and the findings of the published study should be presented instead.  

The report by Elphick et al. on the toxicity of selenium to westslope cutthroat trout (see

https://circle.ubc.ca/handle/2429/24644) should also be included in the literature

review.

iv. The statement that “Observations of fish deformities revealed that most locations had

few or no fish with any observable deformities” is rather misleading.  Selenium induced

developmental deformities occur in the larval stages and those larvae with significant

deformities are unlikely to survive to adulthood.  Thus, the absence of deformed adult

fish does not necessarily mean that selenium is not causing developmental deformities.  

This statement should be removed from the summary of the GEI 2007 report.  However,

the other findings about fish diversity and diversity and density are relevant and

instructive.

4. FIELD MONITORING – EXPOSURE CHARACTERISTICS

A. It is stated in the Report that, “Sulfate has been shown to potentially reduce selenium

bioaccumulation and resulting toxicity.” Presumably this is deemed significant because sulfate

concentrations were found to positively co‐vary with selenium concentrations in the study area

and national selenium standards are not based on studies from high sulfate waters. However, it

is not noted in the Report that sulfate inhibition of selenate uptake, while demonstrated in the

laboratory for environmentally unrealistic conditions (very short‐term exposures without other

chemical species of selenium present), has never been demonstrated to meaningfully alter

bioaccumulation of selenium into aquatic organisms under field conditions, even in selenate‐

dominated waters (e.g., Birkner 1978). The authors of lab studies demonstrating sulfate

reduction of selenium uptake in aquatic organisms clearly recognized the strict limits of their

work for extrapolation to natural waters. For example Hansen et al. (1993:77) wrote, “Thus, at

this time, it does not appear that we have sufficient evidence to justify the consideration of

sulfate as a factor in the regulation of Se in aquatic environments.” Williams et al. (1994:452)

wrote, “At present there is little information available that allows us to assess how relevant this
Page 4

study’s conclusions will be in natural waters containing a complex assemblage of selenium

[chemical] species.” Finally, Ogle and Knight (1996:278) reported that for water concentrations

of selenium near the national criterion of 5 ppb, “... the differences [in selenium

bioaccumulation and toxicity] between extremely different sulfate concentrations are not

significant...” In short, there is not a reasonable scientific basis to expect the sulfate

concentrations in the study area to make any difference with regard to a site‐specific selenium

criterion. [p. 33]

Evaluation:  This is outside my area of expertise.  However, USEPA (2010) states “The

antagonistic relationship between selenium and sulfur, which inhibits selenium

bioaccumulation, is well defined, as sulfate competes with selenate in their uptake into

aquatic organisms (Ogle and Knight 1996; Riedel and Sanders 1996; Bailey et al. 1995;

Hansen et al. 1993; Appendix B).”  USEPA (2010) applied a sulfate “correction” to the acute

toxicity of selenate.

B. Because of the hydrological connectivity between the upstream putative background sampling

sites and the selenium source waters the SSSC Workgroup collectively agreed that a reference

site completely outside the phosphate mining district was also needed. The chosen reference

site was on the South Fork Tincup Creek [p. 14]. The Report later states, “With few exceptions,

concentrations of selenium in both trout and sculpins [sic] tissue from upstream Crow Creek and

Deer Creek locations ( [putative] background) are not significantly different from the reference

location tissue concentrations.” [p. 34].  However, that statement is contradicted by data

presented in Figure 4‐8, Table 5‐2, Appendix A: Table 4 and the discussion of statistical results

on page 6 of Appendix A. Statistical test results presented in Appendix A: Table 4, show that

trout whole body tissue selenium at the reference site is statistically significantly lower than at

the upstream purported background sampling sites on Crow Creek and Deer Creek. This is also

visually evident in Figure 4‐8. Table 5‐2 reports a trout tissue selenium value of 2.56 ug/g dw at

the regionally appropriate reference site outside the mining district, which is typical for fish

from waters unpolluted with selenium (see Walsh et al. 1977; Lowe et al. 1985 and summary in

Skorupa 1998: Table 1). By comparison, fish tissue values in upstream Crow Creek and Deer

Creek, the purported background streams, averaged from about 6‐9 ug/g dw which is well

above the unpolluted fish tissue norms established by national monitoring programs. Trout

movement up and downstream into and out of the exposed stream segments and the purported

background segments constitute the most plausible explanation for these data. Accordingly, the

assumption that the upstream sampling sites represent background conditions uninfluenced by

mining seems unlikely to be valid, especially given the absence of barriers to fish movement in

the study system.

Evaluation:  Partially concur.  The text on page 34 is incorrect and should be updated to

match Appendix A.  The statement that Crow Creek upstream of the confluence with Sage

Creek is unimpacted by mining activities should be clarified to state that surface water and

sediments are unimpacted and that some of the biota there is also unimpacted; i.e., sculpins

(see Appendix A, Table 8), invertebrates (Figure 4‐13, assuming Deer Creek is representative

of unimpacted conditions), and periphyton (Figure 4‐14, assuming Deer Creek is

representative of unimpacted conditions).  The same statistical comparisons that were

performed for fish should be performed for these other environmental media; i.e., the

ANOVA/Kruskall‐wallis test to evaluate whether there are differences in concentrations by

“location” to support these statements.  The interpretive report should acknowledge that

trout in Crow Creek appear to moving back and forth among the “locations” and trout
Page 5

caught upstream of the confluence with Sage Creek may have spent some of their time

downstream of the confluence.  Thus, trout upstream of the confluence may also be

affected by the selenium downstream of the confluence and even in Sage Creek.  It may be

better to focus future work on sculpins, which are probably less vagile, or to install fish

barriers prior to fish sampling.  

C. It should be noted in the Report that brown trout are known to undergo vitellogenesis, the

stage of egg formation when selenium is incorporated into fish eggs, for 6‐7 months prior to

spawning (Estay et al. 2003). This extended period allows ample time for long distance

movement of fish up and downstream. This interpretation of the fish tissue data is further

supported by the results illustrated for periphyton in Figure 4‐14. Periphyton is capable of only

passive movement with water currents and therefore cannot move upstream. In contrast to the

mobile trout, the tissue selenium of periphyton is not significantly different (see also Appendix

A, Table 14) between the regionally appropriate reference site on South Fork Tincup Creek and

the upstream Crow Creek and Deer Creek sites (see also Appendix A, Table 14).

Evaluation:  Concur.  The reference cited indicates that the formation of eggs in trout takes

at least six months.  Combined with the apparent movement of trout in Crow Creek, this

allows an extended period over which egg formation of trout upstream of the confluence of

the Crow and Sage Creeks could be affected by selenium in Sage Creek.

D. In summary, Figure 4‐1 shows tight concordance of water selenium concentrations (< 1‐2 ppb)

for the reference site and the purported background upstream sites, Figure 4‐14 shows tight

concordance of periphyton selenium concentrations (2‐3 ppm) for the reference site and the

upstream sites, yet the trout selenium concentrations are distinctly discordant at the reference

site versus the purported background upstream sites. The much higher trout selenium at the

upstream sites relative to the regionally appropriate reference site would be the unexpected

result, not the low reference site trout selenium, according to current models of selenium

transfer from a periphyton food web base up to fish (see Presser and Luoma 2010). The fact that

the fish sampled at the purported background locations appear to have been influenced by

mining has important implications for the brown trout toxicity study in light of the fact that no

spawning brown trout that could have served as valid field controls were collected from South

Fork Tincup Creek (discussed further below).

Evaluation:  please see response to comments 4.B. and 4.E.

E. The Report indirectly highlights the fundamental problem with population level evaluations of

trout when it is stated, “For this study, trout populations are being evaluated to assess if there

are any obvious [underlining mine] negative impacts to those fish populations due to selenium.”

[p. 40] Clearly, “...obvious negative impacts..” are rarely evident at the threshold levels of

toxicity that are the appropriate focus for establishing water quality criteria; especially in

demographically open populations that can be maintained by immigration. Using fish density

measurements or age‐class profiles as the baseline criteria to establish an effect is indefensible

due to the high baseline variability of such coarse metrics (see review of this topic in Janz et al.

2010: Section 6.7.3). By relying on such coarse metrics the Report fails to clearly distinguish

between measures of habitat attractiveness and habitat quality.  What these investigators

should be looking at in the field are specific demographic parameters that would allow

demographic modeling. For example, Vasterling (2003) took that approach to studying riparian

birds in the Idaho phosphate district and found that even though the species she studied were
Page 6

highly abundant, they were not self‐replacing and that their persistence (at least during the

study years) depended on immigration. She also found evidence for localized selenium toxicity,

despite documenting abundant populations. Low quality polluted habitats that poison fish and

birds can, nonetheless, be highly attractive. For example, Kesterson Reservoir in California was a

site that caused catastrophic levels of reproductive toxicity among birds including complete

(100%) reproductive failure, yet year after year very high densities of breeding water birds could

be found at Kesterson Reservoir, with these demographically open populations being

maintained by immigration of new breeders from outside the system. Yet, if researchers had

gone no further than documenting the high densities of breeding birds at the site they would

have erroneously concluded that no adverse effects on birds were occurring at Kesterson when

in fact this reservoir was a highly dangerous “attractive nuisance” and regionally attracted

waterbirds from other “source” populations into a demographic “sink” population. The trout

populations in the Crow Creek watershed, just like the bird populations at Kesterson, are open

populations with no barriers for movement of individuals into and out of the population.

Accordingly, all of the investigative effort spent and space used in the Report to document trout

densities and draw toxic risk conclusions from those densities [summarized on pp. 46‐47] is

misleading. These data simply show that trout are in the stream just as simple counts of

breeding birds at Kesterson showed birds there were abundant. In neither instance would these

data address the issue of whether the water quality was impaired, or not, from selenium or any

other chemical. This comment also applies to the portion of the literature review that was

devoted to reviewing the many other fish count studies that also erroneously conclude that high

habitat occupancy, simply one measure of habitat attractiveness, equates to high habitat quality

(i.e., low toxic risk). These studies fail to account for the mobility of the species and the

confounding effect of immigration to data interpretation. Studies of that type, as opposed to the

type represented by Vasterling (2003), are simply not designed to rule out the possibility that

they are dealing with contaminant‐induced demographic sink populations that would collapse in

the absence of continual immigration from outside the site. Janz et al. (2010:191) recount a

good case example of this issue. The measured whole body selenium concentrations in

Thompson Creek trout were 4‐14 ug/g dw and fish in this creek persisted across about 20 years

of standardized monitoring. In a nearby tributary and adjacent pond the selenium concentration

in trout averaged 13 ug/g dw whole body, yet the trout in these water bodies collapsed and

disappeared within the first three years of monitoring. The primary difference between

Thompson Creek and the nearby water bodies was Thompson Creek was open to fish movement

into and out of the resident population from other areas, while the other nearby waters were

not.

Evaluation:   

i. Page 36 states “If selenium is having a toxic effect, even on a chronic basis, monitoring

over a period would be expected to indicate some relative reduction in fishery potential,

when other factors are considered.”  As stated by USFWS, immigration may maintain

populations in locations where fitness is reduced due to pollutants.  Therefore, the

statement on page 36 should be amended to take this into account; e.g., “If selenium is

having a toxic effect, even on a chronic basis, fish density/biomass may be reduced, as

long as the effect of selenium toxicity is not masked by immigration from other areas.”

ii. On a scientific basis, I agree with the statement by USFWS that evaluating populations

may not reveal individual level adverse effects when the population evaluated is

maintained by immigration (see, for example Larison et al. 2000).  However, it should be
Page 7

noted that none of the fishes in Crow or Sage Creeks are covered by the Endangered

Species Act and that IDFG (2007) manages Yellowstone cutthroat trout (YCT) on the

population (and drainage) level.  Further, the Clean Water Act and the Consent Order do

not appear to contain a no incidental take provision.  IDFG (2007) states that their goal

for YCT in Crow Creek is to “ensure that mine‐generated selenium is not having

population level effects on cutthroat trout.”  Thus, evaluating the effects of selenium on

the fishes in Sage and Crow Creeks on a population level appears to be in agreement

with the precedent set by IDFG and within the applicable laws.  In contrast, the death of

birds from selenium poisoning at the Kesterson National Wildlife Refuge was a violation

of the Migratory Bird Treaty Act (see Benson et al. 1993).

iii. In order to evaluate whether selenium has had an adverse effect on the fish populations

in Sage and Crow Creeks, additional statistics should be performed, as follows:

a. If there is sufficient historical data, statistical tests should be performed to

determine whether the populations are increasing or decreasing in density

b. The regressions in Figures 4‐24a‐d should be run as either a principal

component analysis (or other ordination technique) or a step‐wise multiple

regression.  When performed that way, the analyses will tell us the relative

contribution to the pattern in fish density and diversity due to selenium

concentrations, flow rate, summer water temperature, stream gradient, and

any other parameters that the authors wish to evaluate (e.g., benthic

macroinvertebrate density/diversity, etc).  Please also provide p values for all

regressions in the figures.  An r

2 by itself is not meaningful.

5. LABORATORY STUDIES – EFFECTS CHARACTERISTICS

A. There are two fundamental concerns with the interpretation of the brown trout toxicity study:

i. The first concern is the lack of any true field control fish from which to calibrate point

estimates of effects. The authors themselves state, “For the brown trout studies, the

response observed at [hatchery] controls was adjusted to the response observed at

background since true controls for this study were not practical.” [p. 50] This is

problematic given the evidence (discussed above) that the upstream Crow Creek and

Deer Creek fish didn’t represent normal or background exposures to selenium, but

probably something on the order of about 2‐times to 4‐times normal background

exposures of 2‐3 ug/g dw whole body fish tissue as found at the appropriate reference

site. Until spawning fish can be collected from South Fork Tin Cup Creek or a comparably

mining‐independent reference site, with whole body tissue selenium < 3 ug/g dw and

egg selenium < 4 ug/g dw (see Table 5‐2 for South Fork Tin Cup Creek data), the point

estimates of toxic effects cannot be confirmed to have been based on proper controls. If

the fish sampled at upstream study area sites are compromised controls, then toxicity

point estimates such as EC‐10s or EC‐20s will systematically overestimate the tissue

concentration of selenium that causes a 10% or 20% toxic effect. To check for this

potential source of systematic bias, comparison toxicity response curves could be

calculated based on the hatchery controls where the data for hatchery fish seem

suitable or on theoretical full performance values where the hatchery control data are

problematic.

Evaluation: USFWS states that the lack of a “true field control” for the brown trout

reproduction study is problematic.  However, it should be noted that there are three

different factors that could be controlled:  1) background selenium exposure levels,

2) maternal fish condition, and 3) parental genetic effects.  The fish from the
Page 8

hatchery used in the brown trout experiment are appropriate controls in terms of

selenium exposures as the selenium concentration in eggs from the hatchery fish

was 0.73 mg/kg‐dw (see Table 6 in Appendix F).  Unfortunately, the figures in

Appendix F are unclear as to whether the length (Figure 3), weight (Figure 4), and

number of eggs (Figure 8) shown for the “hatchery fish” are from the fish that were

used to calculate the ECs (i.e.,  SPC) as opposed to those hatchery fish that were not

used to calculate the ECs (i.e., SC).  However, what figures are presented in

Appendix F do show that there is a difference in fish condition between the wild

caught fish and the “hatchery fish” that might have affected the results.  With

respect to parental genetic effects, IDFG (2007) and Cegelski et al. (2006) show that

the creeks that are genetically closest to Crow Creek for YCT (which I assume would

be similar for brown trout) are Horse and West Pine, which would be the most

appropriate sources for field controls, with the exception of a location somewhere

on Crow Creek.  South Fork Tin Cup Creek is somewhat removed genetically, but still

may be a viable field control in terms of parental genetics, although that question

would best be addressed by the geneticists at IDFG.  Overall, I concur with USFWS

that the lack of a true field control for the brown trout study (Appendix F) adds

some uncertainty to the ECs derived for brown trout and that it would be preferable

to use either fish a) only from the field or b) only from hatcheries.  Mixing the two

adds an additional variable that cannot be controlled without further replication.

ii. The second concern is that the calculated point estimates of toxic effects are not based

on the most sensitive reproductive endpoint. The point estimates are based on the

endpoints of alevin survival to 15 days post‐swim‐up and sum total normal fry across

four categories of deformity/abnormality. The EC‐20s for those two endpoints are 21.6

ug/g and 21.7 ug/g egg selenium respectively (Appendix F: Table 8), yet Appendix F,

Figures 12 and 13, show that the EC‐100 for swim‐up occurs somewhere between about

20 ug/g and 27 ug/g egg selenium. Although many of the brown trout alevins that had

failed to swim‐up by the time this study was terminated may have still been living (2.4‐

33%), in nature total failure of swim‐up would be equivalent to total reproductive

failure. Analysis of the swim‐up endpoint for less severe point estimates of effect than

the EC‐100 should be explored.

Evaluation:  USFWS states that “the EC‐100 for swim‐up occurs somewhere

between about 20 ug/g and 27 ug/g egg selenium” and that the analysis of the

swim‐up end point should be explored.  I used ImageJ to evaluate Figure 13 and

calculated the following:  the first data point for total failure of swim‐up was

approximately 26.7 mg/kg‐dw and the last data point before total failure was

approximately 20.5 mg/kg‐dw.  Thus, the EC10 for failure to reach swim up

would probably be around 20 mg/kg‐dw, which exceeds the lowest EC10 of

17.68 (Table 5‐1).

iii. Related to the reported observation that “... yolk fry never swam up...” (Appendix F:

page 17) for the five sets of eggs with the highest selenium exposures (also recorded as

0 “% Swim‐up” in Table 5‐2 of the Report’s main body and illustrated that way in Figure

12 of Appendix F), it is very perplexing as to how the 15‐day Post Swim‐up feeding trials

produced any data for those same sets of eggs (reported in Appendix F, Figure 37)?

Clearly, before a 15‐day Post Swim‐up Feeding Trial can be conducted alevins have to
Page 9

actually swim‐up. If “... yolk fry never swam up...”, how is Post swim‐up data presented

in Figure 37 generated?

Evaluation:  Concur.  Table 5‐2 shows that the following fish had 0% swim‐up

but were assessed in the post swim‐up survival test: LSV2C‐003, LSV2C‐004,

LSV2C‐005, LSV2C‐010, LSV2C‐021.  This agrees with the number and

approximate concentrations of eggs that did not achieve swim‐up shown in

Appendix F Figure 13.  

iv. The formulation of the “sum total normal fry” endpoint across four

deformity/abnormality categories is not an environmentally realistic way to formulate

that data [Appendix F, page 30 and Figure 44]. For example suppose five fry are each

assessed for the four categories of deformity/abnormality (as per assessment methods,

Appendix F, page 9) and the four effect categories are viewed once each, and each in a

different one of four fry, leaving one fry that exhibits none of the deformity/abnormality

effects. Then the “average” number of normal fry across the four categories is going to

be 80%, 4 out of 5 fry will be free of any particular category of deformity/abnormality

even though only one of the five fry in question, or 20%, will be a fully normal fry. That’s

a large bias in the direction of underestimating toxic effect (and overestimating EC‐10

and EC‐20 concentrations) because environmental realism dictates that any fry not fully

free of deformity/abnormality is unlikely to survive in nature. The deformity data ought

to be re‐analyzed based on a “proportion fully normal fry” endpoint. Such an endpoint

would be expressed on a 0 to 1 scale rather than a 0 to 4 scale.

Evaluation:  Concur.  The proportion of deformed fry should be calculated on a

per fry basis, with each fry assigned a 0 for normal and a 1 for deformed.  Thus,

the “fraction normal” should range from 0 to 1.   

v. When there is no causal relationship between two variables, as is the case with whole

body selenium and egg selenium (dietary selenium is the causal variable for both), the

proper procedure (e.g., see Gill 1987) to back translate from EC‐10 and EC‐20 egg

selenium concentrations to associated whole body selenium concentrations in trout

would be to directly calculate a new regression equation by reversing the original X and

Y axes rather than by algebraically re‐arranging the original regression equation to solve

for whole body selenium as a function of egg selenium as the Report does on p. 65 of

the main text and on p. 33 of Appendix F. Accordingly, the whole body translations from

EC point estimates for egg selenium presented in the Report and in Appendix F warrant

re‐calculation.

Evaluation:  Concur.  As long as the whole body selenium concentration data is

available, the ECs should be calculated directly (e.g., in a regression of whole

body concentration vs. fraction normal fry).  This improves the accuracy of the

calculated result (i.e., unless the r

2 for all equations is 1.0, there is some loss of

accuracy).

vi. The species sensitivity distribution (SSD) illustrated in Figure 6‐1 of the Report does not

appear to be scaled properly on the Y‐axis. For example at the brown trout EC‐20 of 21.6

ug/g dw egg selenium 1 of 7 species is affected or a proportion of 0.143, but the brown

trout data point is plotted at about 0.080 on the Y‐axis. Additionally, based on EPA’s
Page 10

current policy for tissue‐based water quality criteria (i.e., Utah proposed site‐specific

tissue‐based criterion for the Great Salt Lake and 2010 draft proposed revision to the

national chronic criterion) Figure 6‐1 should be re‐drafted based on EC‐10 values rather

than EC‐20 values. If it is believed that results for all the other species of fish illustrated

in Figure 6‐1 are valid for comparison to brown trout results from this specific study

area, then typically EPA would employ such data to estimate the 5th percentile value for

the SSD as an acceptably protective chronic criterion for the entire 10‐species fish

community in the Smoky Canyon study area (of which only two species have been

tested). If it is not believed all the results illustrated in Figure 6‐1 would necessarily

apply on a site‐specific basis to the study area then the only site‐specific sensitivity

conclusion that can be reached is that brown trout are more sensitive than Yellowstone

cutthroat trout. Thus, either a 5th percentile value should be calculated for the SSD

presented in Figure 6‐1, or a safety factor should be applied to the brown trout toxicity

results to acknowledge that only two of ten site‐specific species were tested for toxic

sensitivity to selenium.

Evaluation:  this is outside my area of expertise.

6. UNADDRESSED ISSUES

A. Compliance with the Clean Water Act mandates water standards that will provide for the “...

protection and propagation of fish, shellfish, and wildlife... (PL 92‐500, Sec.101(2)). The

Report is silent on all taxa other than fish and aquatic invertebrates. Recent research

(Presser and Luoma 2010) reveals that fish whole body selenium concentrations are

consistently about 1.1 times the selenium concentrations of co‐located aquatic

invertebrates. The Report asserts that the proposed site‐specific fish egg criterion (21.6 ug/g

dw) would translate to about 14 ug/g fish whole body selenium, which, divided by 1.1,

would translate to about 12.7 ug/g dw aquatic invertebrate selenium. Based on Ohlendorf’s

(2003: Figure 17.1) regression of mallard egg hatchability against dietary selenium exposure

of hens the EC‐10 is only 4.9 ug/g dw. Mallard hens, along with the majority of other species

of water birds, rely heavily on aquatic invertebrates during the breeding season in order to

meet the high protein requirements of ovulation. For a dietary exposure of 12.7 ug/g dw

dietary selenium, Ohlendorf’s regression equation predicts a point toxicity estimate

exceeding 85% failure of mallard eggs to hatch. Thus, it seems highly doubtful that the

proposed site‐specific criterion would comply with the Clean Water Act’s mandate to

protect wildlife. As the studies of Skorupa et al. (2002) and Vasterling (2003) have

documented, in both riparian and wetland systems in the phosphate‐mining region of

southeastern Idaho, avian eggs exhibit elevated selenium concentrations commensurate

with the system wide levels of selenium contamination for species of birds across a

spectrum ranging from blackbirds to mallards to geese to sandhill cranes.

Evaluation:  Concur.  Presser and Luoma (2010) also conclude that “based on

typical TTFs [trophic transfer factors] for Se, USEPA’s proposed whole‐body fish

tissue criterion of 7.9mg/g dw (USEPA 2004) would also allow Se concentrations

in aquatic invertebrates that, when eaten by breeding waterbirds, would pose a

substantively higher hazard (see, e.g., Ohlendorf 2003; EC50) for avian toxicity

than the designed level of protection for fish...”

B. Employing brown trout as a sensitive surrogate for all the untested species of fish in Crow

Creek, the goal would be to keep the selenium concentrations in eggs of all species below
Page 11

the brown trout‐based site‐specific criteria of 21.6 ug/g dw (for the moment neglecting

whether this particular value is defensible or not, see recalculations below). This would be

straight forward to monitor and manage if compliance monitoring were going to be based

on direct measures of selenium in fish eggs. But for practical reasons (difficulty of sampling

ripe female fish, as for example at the reference site) it is actually whole body selenium that

is likely to be monitored. This raises the question of what whole body value would generally

keep egg selenium below the target value in most (if not all) local species of fish?

Determining this is not as simple as merely answering what whole body selenium

concentration is likely to keep brown trout eggs below the target value and then

extrapolating that to all other species of fish, as is done in the Report. The Report is silent on

the rationale for this. Just as species of fish vary in their toxic sensitivity to selenium

exposure, there is also a great deal of interspecies variability in the partitioning of selenium

between different tissues. Studying a community of Colorado River Basin fish species,

Osmundson and Skorupa (2011) found that simple ratios of egg to whole body selenium

vary among species from average ratios of about 1 to 8. The 75th percentile value was about

4. The Report found a ratio for brown trout, at the EC‐20 target for eggs of 1.54, i.e; the eggs

contained about 1.54 times as much selenium as a whole body analysis would yield. Thus, a

proposed egg target of 21.6 ug/g dw “translates” to a whole body target of (21.6)/(1.54) =

14 ug/g dw.  However, a ratio of 1.54 is on the low end of the range of species variability for

such ratios documented for 12 species of fish in the Colorado River Basin. If species

variability of ratios is similar for the 10 species of fish in the Smoky Canyon study area, then

a whole body target of 14 ug/g could result in most species of fish (60‐70% in the Colorado

River Basin data set) actually exceeding the 21.6 ug/g egg target (some species by a lot). The

way to protect against this potential problem is to develop a site‐specific set of tissue ratios

for the 10 species of local fish (or most of them) and then select an appropriately protective

translation ratio from that data set. Presumably such a local data set could be obtained

within a spawning season or two of field sampling, as was achieved for the study of the

Colorado River Basin fish community reported in Osmundson and Skorupa (2011).

Alternatively, data from studies such as Osmundson and Skorupa (2011) could be employed

to estimate a conservative ratio, for example Osmundson and Skorupa’s 75th percentile

value. In any case, arbitrarily extrapolating the brown trout tissue translation ratio to all

other species of fish is inappropriate and could lead to substantive loss of the presumed

protectiveness of a site‐specific criterion.

Evaluation:  Concur.  However, if using ratios from fishes other than those at the

site, it would be more protective to use the 95th percentile, rather than the 75th.

7. RECOMMENDED RECALCULATIONS

I. Hatch to 15‐days post swim‐up fry survival endpoint

a) This toxicity response endpoint for the brown trout toxicity study (Appendix F of the

Final Report) is based on raw data presented in Appendix A of the brown trout toxicity

study (i.e., Appendix A of Appendix F) in Tables 3‐3 and 3‐4. The specific endpoint is the

“Total Survival %” from Table 3‐4 divided by the “% Hatch” from Table 3‐3. For example,

for control sample SPC‐006 these figures would be [(96.0) / (99.7)] for a survival

proportion of 0.963 from hatch until test termination at 15‐days post swim‐up. For

reasons that this reviewer was unable to determine, 27 of the 36 reported survival

proportions for this endpoint reported in Table 3‐4 are incorrect. Of the 27 incorrect

values, 24 are incorrect high and 3 are incorrect low. In some cases the discrepancy is
Page 12

quite substantial. For example, for sample LSV2C‐004 the reported endpoint survival

proportion is 0.662 (66.2%) even though the reported “Total Survival %” in Table 3‐4 of

16.8 and reported “% Hatch” from Table 3‐3 of 50.6 actually yield a “hatch to

termination” endpoint survival proportion of 0.332 (33.2%; i.e., 16.8 / 50.6). Thus, to

begin with for this recommended recalculation the 27 incorrectly reported results were

corrected.

Evaluation: The tables in the Interpretive Report and its appendices have

headings that do not accurately reflect the data in those tables.  For example,

one of the column headings is “%hatch,” however, this is defined in the text of

Appendix A of Appendix F page 2‐10 “as the number of live fish and alevins at

day of first hatch compared to the number of eggs at test initiation.” Thus, the

actual number of eggs that hatched should be higher than the %hatch shown in

the tables.  As defined this way, I am unsure of the utility of “%hatch.” Further,

if additional eggs hatched after the day of first hatch, USFWS’ re‐calculation is

invalid and the data cannot be reliably evaluated until all endpoints are clearly

defined and the actual raw data is presented in terms of the number of

individuals surviving at each stage.

Below, I list the endpoints from Tables 3‐1, 3‐3 and 3‐4 and indicate which

endpoints are unclearly defined in the Interpretive Report:

1. #Eggs placed in study: self‐explanatory

2. Day of first hatch: self‐explanatory

3. %hatch

a. Defined as “Percent hatch was determined as the number of live

fish and alevins at day of first hatch compared to the number of

eggs at test initiation.”  Therefore, this column in the tables should

be labeled something like “%hatch on 1st day”

b. a second column should be added showing the total number of

individuals that hatched and labeled “%hatch”

4. Day of swim up: please define more completely.  Is this the day that the first

individual reached swim‐up or the day that some % reached swim‐up?

5. % Swim‐up: the number of individuals that reached the swim‐up stage

6. Survival (%) at Swim‐up Stage:  this appears to be defined as the number of

individuals alive when some % of those individuals have reached swim‐up,

including individuals that have not reached swim‐up.  However, I do not

think that this variable is useful and should be deleted.

7. Survival (%) in 15‐d Post swim‐up stage:   

a. This variable is self‐explanatory

b. However, in order to be transparent, the tables need another

column to explain this variable; i.e., “# in 15‐d Post swim‐up test”

8. Total Survival (%): it is unclear how this variable has been defined.  Using SC‐

001 as an example, if there were 600 eggs and 22.8% reached swim‐up,

then 137 individuals reached swim‐up.  If 100 of those individuals were then

placed in the 15‐d post  swim‐up test and 97 survived, to determine the

number that are assumed to have survived to 15‐d after swim‐up, one

would multiply 97% by 137, to get 133.  However, 133/600 is 22.2% and not

the 22.7% shown in Table 3‐4.  Please define how this was calculated.  
Page 13

Please also note that this is an estimated number and not an actual

observation.

9. Survival (%) from Hatch until test term.: self explanatory.  But, it should also

be noted that this is an estimated number and not an actual observation.

I have translated these variables into number of individuals on the following page for SC‐001 and CC‐

150‐009 to show what the raw data tables should look like.  Then, the percentages can be calculated

from this in a manner that is completely transparent to the reviewers.  Note that because some of the

data was not provided, there are some “?” in my example tables.
Page 14

Suggest raw data format:

Variable 1 2 3a 3b 4 5 7b 7a 

Treatment

#Eggs

placed in

study

Day

of 1st

hatch

#hatch

on 1st

day

#eggs

hatched

Day of

swim

up

#

reached

swim‐

up

# entered

into post

swim‐up test

# alive at

15d post

swim‐up

SC‐001 600 42 143 ? 69 137 100 97 

CC‐150‐009 600 39 171 ? 72 163 100 99 

Notes: * calculated as 5*(7a/7b)

Suggested table format for presenting percentages

Calculation 3a/1 3b/1 5/1 7a/7b 8/1 8/3b

Treatment

%hatch

on 1st

day

%eggs

hatched

%

reached

swim‐

up

Survival (%)

in 15‐d Post

swim‐up

stage

Estimated

total

survival

(%)

Survival (%)

from Hatch

until test

term.

SC‐001 23.8 ? 22.8 97 22.2 ?

CC‐150‐009 28.5 ? 27.2 99 26.8 ?
Page 15

b) Next, there were five samples (LSV2C‐003, LSV2C‐004, LSV2C‐005, LSV2C‐010, and

LSV2C‐21) that did not produce any swim‐up fry by the time the experiment was

terminated. Instead of recording a post swim‐up fry survival of 0.0% for those samples,

the authors of the final report chose to record the percent survival among the

developmentally “dead‐ended” alevins. By doing this, they are assessing a post swim‐up

fry endpoint for most of the samples and a developmentally completely different dead‐

end alevin endpoint for the five samples mentioned above. Mixing two different

endpoints for exposure‐response modeling fundamentally invalidates the modeling. This

error is easily corrected by using only the post swim‐up fry endpoint and recording a

post swim‐up fry survival of 0.0% for the five samples in question.

Evaluation:  Concur.

c) The authors excluded data from the hatchery control samples because the first set of

hatchery controls performed very poorly with regard to % hatch and the second set of

hatchery controls were obtained as eyed embryos and therefore weren’t comparable to

the starting point for all the other samples (started as undeveloped fertilized eggs).

While the second set of controls yielded excellent % hatch performance, that may have

been due to the advanced stage of the embryos at receipt and thus much shorter

holding time to hatch (9‐13 days) than for any of the other samples (which had holding

times to hatch of 37‐47 days). In addition to those data exclusions, there were 10

additional samples with missing test organisms because of clogged drain pipes that

caused water to overflow out of the test system, apparently carrying some test

organisms away (see Appendix A of Appendix F). The authors do not know how many of

the carried away organisms were dead or alive, deformed or normal, or anything else

about them. The authors chose to simply ignore the missing test organisms, effectively

assuming the proportion of missing fish that were live or dead and deformed or normal

was exactly the same as those proportions in the fish left behind. There is no valid

justification for such an assumption. Best scientific protocol dictates that the uncertain

results for these 10 samples also be excluded from exposure‐response modeling.

Evaluation:  Concur.

d) Thus, the valid data set for modeling the “hatch to15‐days post swim‐up fry survival”

endpoint is as follows (Table 1; log transformation of exposure data is performed on ppb

exposure instead of ppm exposure to avoid negative logarithms): [table not shown here]

Evaluation:  as presented in above, I think that USFWS is assuming that the

definition of “%hatch” is the total number of eggs that hatched, instead of the

number of eggs that hatched on the first day of hatching (which was used in the

Interpretive Report).  Therefore, I do not think that the calculated numbers

presented by USFWS are correct.  The authors of the Interpretive Report should

weigh in on this and provide tables as I have suggested above for re‐evaluation

by USFWS.  

e) The authors of the Report employ EPA’s Toxicity Relationship Analysis Program (TRAP)

for modeling exposure‐response. However, TRAP offers several options with regard to

statistical models. The authors chose to use the Least‐Squares Nonlinear Logistic
Page 16

Regression model (see Appendix E of Appendix F). EPA advises in the TRAP program

overview that the Least Squares Logistic Regression model is most appropriate for

endpoints that are (1) measured on a continuous scale (for example weight or height)

and (2) expected to have infinite tails approaching 0% and 100% response. Neither of

these conditions applies to the brown trout data in question. The data in question are

(1) non‐continuous dichotomous (binomial) data (each hatched egg either goes on to

become a 15‐day post swim‐up fry or does not) and (2) the endpoint would be expected

to have a finite “threshold” response in which there is no response until a minimum

finite threshold exposure is experienced, not an infinitely gradual response tail. For

dichotomous, threshold response data, EPA advises that the most appropriate TRAP

model choice is the Tolerance Distribution Analysis fitted to a Triangular Distribution.

Therefore, for this recalculation those are the modeling options that were chosen for

TRAP exposure‐response modeling with the following result (Figure 1): [figure not

shown here]

Evaluation:  Concur, TRAP does recommend a triangular distribution for

dichotomous data.  However, I would suggest that USEPA’s BMDS (see

http://www.epa.gov/ncea/bmds/) or Cetis (https://www.tidepool‐

scientific.com/index.htm) be used instead.  These other programs have several

more dichotomous models than TRAP which will allow the authors to determine

which model best fits the data instead of assuming that a logistic model is the

best fit.

f) The lack of useable data points within the intermediate response range violates the

TRAP program requirements for model fitting, generating an error warning message.

The EC‐10 and EC‐20 estimates of 14.9 and 16 ppm shown above are the result of

forcing a model fit and should not be viewed as reliable. Recalculation for this endpoint

reveals that the useable dataset is not sufficient for reliably estimating any EC‐10 or EC‐

20 values. The only reasonably certain information that these data yield is that the finite

threshold for total response, the EC‐100 threshold, occurs somewhere between an egg

selenium concentration of 14 ppm and 27 ppm. Because 100% response (i.e., total

failure of brown trout alevins to achieve swim‐up) is already being expressed at an egg

selenium concentration of 27 ppm, and because valid estimates of the EC‐10 and EC‐20

cannot be modeled from the useable dataset, proposing an egg selenium standard of

nearly 22 ppm (as the authors do in the final report) based on an invalid EC‐20

estimation (as the authors do in the Report) does not seem very defensible. Any site‐

specific standard proposal higher than 14 ppm egg selenium cannot be defensibly

supported by the “hatch to 15‐day post swim‐up fry survival” endpoint.

Evaluation: As indicated above, I think that the authors have defined “%hatch”

in a different manner than assumed by USFWS and, therefore, any re‐

calculations based on the % or # of individuals that hatched should wait until the

authors clarify their endpoints and provide the raw (i.e., number of individuals)

data.  However, if USFWS is correct in their calculations, there is a data gap in

that there are no data points between almost total survival and almost total

mortality.  In that case, almost all dichotomous models should fit the data and

the EC10 or EC20 will vary depending on what model is chosen to fit the data.

II. Larval Deformity Endpoint
Page 17

The collection of larval deformity data was seriously flawed in at least two ways.

a) First, the larvae assessed for deformities were not a random subsample of all larvae.

Alevins and post swim‐up fry that died during the brown trout toxicity study mostly did

not get assessed for presence or absence of deformities (see Section 2.5 of Appendix A

to Appendix F). Since deformed larval fish can be expected to have a differentially high

propensity to die compared to undeformed larval fish, all the measures of percent

deformed larvae are likely biased low. For example, among a limited number of test

groups (samples LSV2C‐004, LSV2C‐005, and LSV2C‐021) deformity assessments are

available for reasonable samples of both dead and live larval fish as follows (Table 2

from raw data presented in the first three of the un‐numbered Tables in Appendix C of

Appendix F): [table not shown here] As the above comparisons show, the bias

introduced by assessing mostly live (surviving) larval fish for deformities can be

substantive. However, the comparisons shown above are for test samples that had high

selenium exposures (27‐36 ppm egg selenium) and presumably therefore would express

this type of bias to a greater extent than test samples from lower exposures would.

Presumably, lower exposures would not uniformly produce exclusively (100%) deformed

dead larval fish. Since the authors did assess greater numbers of dead fish for the five

highest exposure groups and the lower exposure groups often yielded predominantly

surviving larval fish, it is possible that the overall magnitude of this bias on exposure‐

response modeling for larval deformities is relatively modest. Nonetheless, the main

points to consider are that the exact amount of bias is unknown, and whatever the

magnitude of bias is it is certain to be in only one direction, underestimation of

response values (thus overestimation of toxic threshold concentrations).

Evaluation:  the table provided by USFWS shows the percentage of deformed

live and dead fry.  The number of deformed dead fry was taken from the first

un‐numbered table in Appendix C of Appendix F.  However, the number of

deformed live fry, as presented in the table generated by USFWS, is not

presented in Appendix C of Appendix F.  USFWS appears to have generated this

number by subtracting the number of deformed dead fry in the first table in

Appendix C of Appendix F from the number of fish from the second table in

Appendix C of Appendix F.  The second table does not indicate whether it is

displaying the results for live, dead, or live and dead fry.  Some values in the

second table are also underlined and some sample IDs are highlighted, which

are not explained.  If USFWS is correct in their assumption that the second un‐

numbered table is for both live and dead fry, then there were no craniofacial

deformities in the live fry for LSV2C‐004; i.e., the number of deformed fry in the

second table is 6+80+15=101, which is the same as is shown in the first table.  

Nonetheless, if USFWS assumption is correct that the second table shows both

live and dead fry, then dead fry may be expected to have higher deformity rates

than live fry and the loss of some dead fry due to decomposition etc. can be

expected to lead to an under‐estimate of deformity rates for those samples that

lost dead fish due to decomposition.

b) Second, each of the four types of deformity examined were assessed and reported

completely independently of each other (see Appendix C of Appendix F). Thus, while it is

possible to determine what percent of assessed larval fish were normal for each type of

deformity, it is not possible to determine what percent of assessed larval fish were
Page 18

completely undeformed (i.e., were fully “normal” larval fish). Unfortunately, the raw

deformity assessment data were not available for review, but were cited (see p. 3‐10 of

Appendix A of Appendix F) as being presented in a separate document (never made

available to this reviewer). The authors of the Report used an endpoint that they called

“sum total normal fry” which in reality should more accurately have been called “sum

partial normal fry” because it is unknown if any of the fry in this sum are simultaneously

normal for all four types of deformity. It is only known that the fry in this sum are

normal for at least one of the four types of deformity (craniofacial, vertebral, fin,

edema). Of course the sum of partial normal fry will necessarily overestimate the sum of

fully normal fry and it is only fully normal fry that can be expected in nature to be free of

any deformity‐related toxic impairment. A recalculation that can be done here to

partially correct for the overestimation of fully normal fry is to recognize that the

proportion of fully normal fish cannot have been any higher than the lowest proportion

of normal fish in any of the four deformity categories reported in Appendix C of

Appendix F. This will be referred to as “maximum fully normal fry”. That endpoint

should be a much better approximation of percent fully normal fry than the author’s

“sum partial normal fry” endpoint. Also, for this recalculation the second set of hatchery

controls provide useable estimates of expected background deformity rates because the

hatching rates were all acceptable control values (i.e., >90% hatch) and from hatch to

15‐days post swim‐up fry these controls are strictly comparable in their holding times to

the field‐exposed test groups. However, the results from the “accidental overflow” test

samples must again be excluded from analysis. This yields the following data set (Table

3) for modeling the response variable of larval deformity (log transformation of

exposure data is performed on ppb exposure instead of ppm exposure to avoid negative

logarithms): [table not shown here]

Evaluation:  Concur.  I checked the first two entries on the table generated by

USFWS and found them to be correct.  

c) This is a repeat of Comment 7.I.e above.  Please see response above.

d) This dataset met all TRAP program requirements for a clean model fit without any error

warnings. The estimated EC‐10 is 6.5 ppm egg selenium and the estimated EC‐20 is 9.0

ppm egg selenium. These estimates are not fully corrected for the “sum partial normal

fry” bias compared to proportion fully normal fry and therefore are still overestimates

of the EC‐10 and EC‐20 for the larval deformity endpoint. Even so, both of these

recalculated estimates are substantially lower than the site‐specific standard of almost

22 ppm egg selenium proposed by the authors of the Report. This endpoint also

illustrates why it is so crucial to eventually obtain valid field control data. If the hatchery

control data points are omitted, and it is erroneously assumed that the CC‐150 and CC‐

350 (upstream) sampling sites are valid field control sites then the estimates of the EC‐

10 and EC‐20 for larval deformity respectively increase to 15.2 and 17.5 ppm egg

selenium as illustrated below (Figure 3): [figure not shown here] Obtaining valid field

control results for the larval deformity endpoint should be a very high priority for future

research.

Evaluation:  Concur.  IDFG should also be consulted to determine which Creeks

have the most genetically similar fish to avoid potential population‐specific

differences in the responses to selenium.
Page 19

III. Regressing X on Y for Estimation of Critical Whole Body Tissue Concentrations. Because

of the distinctly species‐specific nature of selenium partitioning between tissue types

within individual fish (Osmundson and Skorupa 2011) and because of the complex

spatio‐temporal biology of selenium deposition into brown trout eggs which can be

expected to differ in a hatchery versus in the field (Estay et al. 2003), this recalculation is

based only on the field data for brown trout as presented in Table 5‐2 of the final

report. Because tissue concentrations from polluted environments typically conform to

lognormal distributions, the regression recalculated here is calculated on log‐log

transformed data. The recalculated regression is illustrated below (Figure 4). The

regression is highly significant (p < 0.0000001) and the adjusted R‐square value for the

regression is 0.788.  This regression yields a critical whole body selenium concentration

of 12.32 ppm at an egg selenium concentration of 21.63 ppm, a result that is slightly

lower than the value of 13.35 ppm critical whole body selenium reported for brown

trout data in the Report based on algebraic re‐arrangement of the egg selenium on

whole body selenium regression equation (see p. 65). When the recalculated regression

above (Figure 4) is evaluated at the EC‐10 for maximum fully normal fry (Figure 2) of 6.5

ppm egg selenium it yields an estimated critical whole body selenium concentration of

5.5 ppm.

Evaluation:  I tested the distributions of the brown trout whole body and

selenium concentration data collected from Sage and Crow Creeks shown in

Table 5‐2 using USEPA’s ProUCL v4.1.0.1

(http://www.epa.gov/osp/hstl/tsc/software.htm).  Neither data set fit a normal

distribution, but both fit a log‐normal distribution, as assumed by USFWS.  I also

re‐calculated the regression presented by USFWS and got the same result.  

Using this equation, 6.5 ppm selenium in eggs does indeed convert to 5.5 ppm

in whole body.  Converting the proposed site‐specific criterion (see Page 71) of

21.63 mg/kg‐dw in eggs using the equation derived by USFWS yields 12.33

mg/kg‐dw whole body whereas converting the site‐specific criterion using the

equation on page 65 yields 14.0 mg/kg‐dw whole body.  Overall, I concur that it

is better to re‐calculate the regression from the raw data rather than to

algebraically re‐arrange the equation.

IV. Test of Assumption that the 14 ppm Proposed Critical Whole Body Tissue Concentration

Derived from Brown Trout Tissue Partitioning Data Would Generally Keep Egg Selenium

Values Below a Critical Value of 21.63 ppm Across a Multi‐species Community of Fish.  

Tissue partitioning ratios from elevated selenium exposure can be expected to be dose‐

dependent. Therefore to examine this issue using the fish community dataset from the

Colorado River Basin reported in Appendices 1 and 2 in Osmundson and Skorupa (2011),

only the fish samples with egg selenium results relatively near (in the range of 15 to 30

ppm) the target value of 21.63 ppm proposed in the Report were evaluated. The results

are presented below (Table 4): [table not shown here].  Thus, for the Colorado River

Basin community of fish, assuming that a critical whole body concentration of selenium

derived from brown trout data would be universally applicable would have allowed

other species of fish to reach egg selenium concentrations as high as 138 ppm, or more

than 6‐times the proposed site‐specific standard of 21.63 ppm. Moreover, the overall

exceedance rate would be 68% (weighting equally each species exceedance rate in Table

4). In other words, instead of the 80% protection afforded to brown trout by an EC‐20
Page 20

standard, the protection afforded to the rest of a multi‐species fish community in this

case would be only 32%. Consequently, in this case, assuming that all fish species

partition selenium between tissues comparably to brown trout would be a very under

protective assumption. This exercise does not prove that such an assumption would also

be very under protective as applied to the Crow Creek fish community, but it does

emphasize that making such an assumption for the Crow Creek fish community (as is

done in the Report) would seem unlikely to hold true and therefore must be validated

before it can be accepted for regulatory purposes.

Evaluation:  On page 71, the Interpretive Report proposes a site‐specific

criterion of 21.63 mg/kg dw egg selenium based on brown trout.  A whole‐body

criterion is not proposed.  Using the equation proposed in the Interpretive

Report on page 65, this would convert to a whole body concentration of 14.0

mg/kg while using the equation derived by USFWS, this would convert to a

concentration of 12.33.  Thus, USFWS is correct in comparing expected whole

body concentrations in fish from the Colorado River to 14 mg/kg‐dw.  I also

checked USFWS calculations in the table they provided.  There was some

rounding in the table presented by USFWS that was not present in the original

report but I matched all but one of the expected whole body concentrations

using USFWS numbers.  Thus, USFWS is correct that an egg concentration of

21.63 mg/kg‐dw may allow some species to have whole body concentrations

over 14 mg/kg‐dw.  Overall, I concur that a selenium criterion of 21.63 mg/kg‐

dw in eggs may not be protective of the other species of fish in Crow and Sage

Creeks.

References

Benson SM, Delamore M, Hoffman S.  1993.  Kesterson Crisis.  Journal of Irrigation and Drainage

Engineering 119:  471‐483.

Cegelski CC, Campbell MR, Meyer KA, Powell MS.  2006.  Multiscale genetic structure of Yellowstone

Cutthroat Trout in the Upper Snake River Basin.  Transactions of the American Fisheries Society

135:711–726.

Idaho Department of Fish and Game.  2007.  Management Plan for Conservation of Yellowstone

Cutthroat Trout in Idaho.  Available at fishandgame.idaho.gov/public/fish/planYellowCutthroat.pdf

Larison JR, Likens GE, Fitzpatrick JW ,Crock JG.  2000.  Cadmium toxicity among wildlife in the Colorado

Rocky Mountains.  Nature 406:181‐183.

Presser TS and Luoma SN. 2010. A methodology for ecosystem‐scale modeling of selenium. Integrated

Environmental Assessment and Management 6:685‐710.

United States Environmental Protection Agency (USEPA). 2008.  Effect of selenium on juvenile bluegill

sunfish at reduced temperature. EPA‐822‐R‐08‐020.

United States Environmental Protection Agency (USEPA). 2010.  Draft aquatic life water quality criteria

for selenium – 2010.  EPA‐822‐R‐10‐001. Office of Water, Washington, DC.
1

Peer Review of the FWS Technical Review Document based on the report entitled: 

Interpretive Findings for Field and Laboratory Studies and Literature Review in Support of a 

Site-Specific Selenium Criterion, Smoky Canyon Mine (August 2010)

Peer reviewer: David M. Janz, Ph.D

Date: October 25, 2011

Preamble: I am a Professor of Veterinary Biomedical Sciences and Acting Director of the 

Toxicology Centre, University of Saskatchewan, Canada. I hold a PhD in Pharmacology and 

Toxicology, a MSc in Environmental Resource Studies and a BSc in Ecology. My main area of 

research is aquatic ecotoxicology, with a focus on developmental and reproductive toxicities in 

fish. I have conducted field and laboratory research with over 20 species of fish, including four 

salmonid species, over the past 25 years. I have published 76 peer-reviewed scientific articles, 

including 22 on the aquatic toxicology of selenium in fish. I should also state that since 2003 I 

have been Principal Investigator on three large grants investigating the aquatic ecotoxicology of 

selenium that were sponsored by the mining industry; thus I am cognizant of the regulatory 

issues and “sensitivities” associated with this trace metalloid. Overall, I feel very well qualified 

to provide this peer review on the FWS Technical Document. My goal from the outset is to 

provide a balanced review based on the principles of sound science. In addition, I should state 

that I have no conflict of interest in any way related to this review.

The Smoky Canyon Mine report provides extensive documentation in support of their 

application to establish an interim site-specific Se criterion to protect aquatic-dependent life 

downgradient of the mine, specifically Se-laden groundwater discharge into a series of creeks 

that support fish communities. Significant financial resources have been devoted to the field 

and laboratory research conducted to develop this report, and as a scientist who does similar 

work I acknowledge the effort required to produce this report.

Selenium is known to cause a spectrum of adverse developmental effects in early life stages of 

fish, which have the potential to impair recruitment into fish populations and thus sustainability 

of natural fish populations and communities. The experimental studies found within the Smoky 

Canyon report are thus focused on these toxic responses in two resident salmonid fishes, 

brown trout and cutthroat trout. Furthermore, the FWS Technical Document is focused on the 

scientific soundness of these experiments and the resulting data analyses. My peer review will 

proceed through the FWS Technical Document and provide an assessment of the validity, clarity 

and fairness of the criticisms found within, as well as other comments related to the document 

and report. Throughout my review, I will refer to the Smoky Canyon Mine report as the “report”

and the FWS review as the “document”.

1.0 Introduction

Paragraph 1: I agree with this comment and find it strange that only a preliminary analysis of 

the YCT data are provided given the extensive nature of the report. Perhaps there was a 

deadline and these analyses could not be finalized by that time? Given the length of time 
2

between report submission, FWS commentary and this peer review, one would think that by 

now the YCT dataset would be finalized and included. Nevertheless, the FWS document is 

directed solely on the brown trout dataset and thus my comments will focus on that aspect.

Paragraph 2: I agree with this comment, although a criterion based on aqueous Se 

concentration implies aqueous exposure as the primary route of uptake in fish, as is true for 

most other trace elements but not Se. Thus the flaw with an aqueous Se criterion for Se. I agree 

that the report neglects to acknowledge that dietary exposure at Belew’s Lake was the major 

route of exposure, and that this was “integrated” into the 1987 USEPA criterion of 5 μg Se/L.

This can easily be clarified in the report.

Paragraph 3: I agree with the first statement, and will expand this to say that in my opinion the 

literature review provided in the report is superficial in many respects and indicates a relative 

lack of knowledge of the current state of understanding with respect to the aquatic 

ecotoxicology of Se. The report should explicitly acknowledge the differences in experimental 

design between the bluegill and trout studies.

Paragraph 4: I do not completely agree with this comment. Although it was an unfortunate 

oversight in the follow-up study with bluegill that photoperiod was not appropriately 

controlled, water temperature is the major driver of physiological responses in poikilotherms 

such as fish. In my opinion the follow-up study is not irrelevant and in fact certain aspects of 

Lemly’s (1993) study were consistent (i.e., reduced survival in fish at 4°C compared to 9°C) even 

though growth and lipid stores did not differ between temperature treatments. In addition, 

since Lemly (1993) was also a laboratory study, what applicability does it have to free-ranging 

fish? I would suggest “softening” this comment to address the uncertainties in both the 

McIntyre (2008) and Lemly (1993) studies.

2.0 Site Setting

My comments concerning appropriateness of the upstream “background” sampling sites can be 

found below under section 4.0.

3.0 Literature Review

I agree with the FWS reviewer that certain aspects of this section were “uneven”, and the 

example given (Besser and Kennedy studies, and associated critiques) is a good indication of 

this. I also found the literature review to be very superficial, demonstrating limited knowledge 

of Se biogeochemistry and toxicity. I could provide many detailed comments here, but since the 

FWS document does not focus on this, then neither will I.

4.0 Field Monitoring

Paragraph 1: Although the sulfate argument has often been used to downplay the potential 

bioavailability of selenate to food webs, I agree with the reviewer’s comments in the document 

that its relevance in field situations is not established. The only ecologically (field) relevant 

competition between sulfate and selenate would potentially occur during uptake into algae 

(and possibly other primary producers such as bacteria and fungi), although this is not clearly 
3

established. If true, sulfate may reduce the bioavailability of selenate to algae, and thus the 

enrichment of Se in aquatic food webs. However this seems to me a moot point since it is clear 

from the near field study sites that sufficient Se is being assimilated and biotransformed by 

primary producers to cause diagnostic features of toxicosis in resident fish (terata). I agree that 

the presence of elevated sulfate at the study sites has no bearing on developing a site-specific 

criterion.

Selenate can theoretically be taken up across the intestine of fish via a sulfate transporter; 

however this is certainly not relevant for freshwater fish as they do not drink appreciable 

volumes of water (unlike marine teleosts). Furthermore, sulfate transporters have not been 

identified in fish gills, so there is no basis for competition between the sulfate and selenate

anions at the gill-water interface, as determined for certain cations using biotic ligand modeling 

approaches. Very little is known in invertebrates. The most relevant study cited by the FWS 

reviewer is Williams et al. (1994) since it was focused on algae. The Hansen et al. (1993) study is 

of questionable significance since it was conducted in benthic and pelagic invertebrates, which, 

like fish, will obtain the majority of their Se body burden via diet.

Paragraphs 2-5: I agree with the reviewer that there are significant (statistically and 

toxicologically) differences in whole body Se concentrations between the “reference” and 

“background” (i.e., upstream) sites. Indeed the data in the report show this, as indicated by the 

reviewer. The 2.56 μg/g dw value for Tincup Creek fish can be considered “normal”, while the 

6-9 μg/g dw range for Crow Creek and Deer Creek sites is greater than what could be 

considered even “high normal” and in fact is approaching or at a toxicity threshold value for 

certain fish species. 

The paragraph discussing the extended period of brown trout vitellogenesis and lack of 

differences in periphyton Se concentrations between reference and background sites is key to 

this entire section and the argument that brown trout are likely migrating between the 

exposure and background sites within the same open watershed/drainage. I strongly agree with 

these assessments made by the FWS reviewer. The reviewer may also consider including 

mention of recent work by Vince Palace’s group using laser ablation ICP-MS in otoliths to record 

Se exposure chronology in free-ranging trout inhabiting lotic systems in a coal mining region of 

Alberta (Palace et al. 2007).

In reading the report, I was not able to find any acknowledgement of even the possibility of 

trout migrating between exposure and background sites. This should be added to the report. 

In this section, and throughout the FWS review document, I recommend explicitly stating 

“whole body Se” when referring to the report data to avoid confusion (instead of, for example 

“trout tissue selenium value”). I also recommend not using ppm and ppb as units of 

concentration; use μg/g and μg/L, respectively.

Paragraphs 6-9: I strongly agree with the statements provided in this section. The mere 

presence of fish in an aquatic system does not necessarily constitute stable, sustainable 
4

population dynamics. Reduced fish abundance at an impacted site can reduce competition for 

resources (abiotic and biotic) and thereby act as an attractive “sink” for immigration. It should 

also be noted that using electrofishing to obtain CPUE data is problematic among sites with 

differing conductivity, since conductivity has a major influence on the ability to stun fish. A 

combination of classic mark-recapture studies at these sites, ideally in combination with otolith 

microchemistry (LA-ICP-MS, mentioned above) would provide a much more robust way to 

evaluate any “obvious negative impacts to fish populations”.

In section 4.6 page 46 of the report it is stated that benthic invertebrate abundance and 

richness did not differ among sites. It should be noted that this is commonly observed in such 

field studies and is not an indication of “no impact” of selenium, since benthic invertebrates are 

tolerant of high Se exposure. The key question is not what the abundance/diversity of 

invertebrates is, but what the Se concentrations are in these benthic invertebrates (see Table 

13, Appendix A), since they provide the majority of trout dietary items.

5.0 Laboratory Studies

This is an important section of the document and I have read Appendix F and Appendix A of 

Appendix F very carefully.

I will say from the start that I am concerned about the very high and variable embryo mortality 

across all sampling sites and the first hatchery collection. I assume this was due to fungal 

growth as is common when aseptic technique is not followed closely. Of concern is the 

documented use of NaCl and especially formalin to control fungal growth, and how this might 

influence endpoints. Overall, the highly variable mortalities would appear at first glance to 

hamper the ability to determine robust concentration-response relationships. In addition, the 

study suffers from the problem of having most data in the “normal” range (although the 

upstream “background” sites cannot be considered “normal) with relatively few data describing 

the threshold for adverse effects (similar to Holm et al. 2005). However, these are the data that 

were collected, and this is field work after all.

I strongly agree with the first concern brought forth by the reviewer, that using responses 

obtained from upstream “background” site trout, with their relatively high Se body burdens, 

will likely overestimate EC10/EC20 values determined for exposure site trout. I agree that 

hatchery data should be used for comparative purposes, and perhaps a composite of the 

hatchery and upstream site data could be used if possible. Are the hatchery trout of the same 

strain as the wild trout? This needs to be considered before using hatchery data. Ideally, 

obtaining spawning brown trout from a true reference site in this region is the best approach. 

Looking at the maps supplied in the report, there doesn’t seem to be a paucity of potential 

locations to collect such fish. Although not ideal, conducting a study in a separate year using a 

proper reference site would seem to be a good approach, perhaps in combination with 

additional sampling of brown trout from the near field exposure sites to increase the sample 

size within the critical (“threshold”) portion of the concentration-response relationships.
5

With respect to the second concern (page 7 of document), I agree in part but have some 

comments to consider. 

I have a major problem with use of the endpoint “survival to 15 days post-swim-up”, purported 

to be a measure of the ability to transition from endogenous to exogenous feeding. I do not 

doubt the ecological relevance of switching to exogenous food as an important life history trait 

in oviparous fish species. However, to my knowledge this has not been demonstrated in any 

previous studies to be a sensitive endpoint associated with elevated Se exposure. Just because 

the data for this endpoint provide a good “fit” using the dose-response software doesn’t 

necessarily mean that they are toxicologically meaningful. My reasoning for this is found below.

Elevated Se has been shown to alter intermediary metabolism related to lipid and carbohydrate 

homeostasis in mammals (Mueller et al. 2008), which has recently been at least partially 

confirmed in fish (Thomas and Janz 2011; Wiseman et al. 2011). Thus, it is distinctly possible 

that fish with higher Se exposure during development may have altered appetite in response to 

this metabolic phenomenon, such as the fry fed for 15 days following swim-up. If true, this 

would confound the results obtained, since fish with greater Se exposure may eat more readily 

compared to controls. In addition, I could find no details on the methodology for feeding fish 

post-swim-up (e.g., type of food, feeding ration, feeding frequency, removal of uneaten food, 

etc). Feeding studies must be very carefully controlled in order to obtain meaningful data. The 

most relevant type of feeding study would employ live prey (e.g., zooplankton) as food items to 

mimic natural conditions.

I am confused with the reviewer’s assessment of “sum total normal fry” vs. “fraction normal 

fry”. From what I can see the report does appear to use fraction normal, not sum. However 

after spending much time reading through the report, it was not clear to me how exactly this 

was calculated. The reviewer makes a strong case (paragraph 3, page 7, starting with “The 

formulation of the “sum total normal fry”...) for the apparent problems with the manner which 

the deformities were expressed. I agree with this assessment, and it appears that the report 

may not be presenting the data in the proper way. My own rationale for this can be found in 

the following paragraph (which could possibly just be a different way of saying what the 

reviewer states in this section of the review). I think the wording in this section of the FWS 

review could be worked on for clarity.

In my opinion, the analyses for deformities should be performed as follows. The various 

abnormality categories (skeletal, fin, craniofacial and edema) do not usually occur 

simultaneously in fish; usually just one or two is evident except at very high exposures. It can be 

assumed that the presence of just one deformity category will have negative consequences on 

fish survival in the wild. As stated by the reviewer, “environmental realism dictates that any fry 

not fully free of deformity/abnormality is unlikely to survive in nature”. I strongly agree with 

this statement. Thus, in my opinion data should be evaluated as simple presence or absence 

(frequency) of any specific deformity; if any deformity was present, that individual fry would be 

assessed as not normal. As far as edema, the report states that this can be reversible and may 

not influence survival, as stated in Appendix C of Appendix F (“Edema was not originally 
6

scheduled for assessment because it was thought sometimes not a teratogenic effect and may 

be transitory as fish develop”). I personally tend to disagree with this given the severity of 

edema often observed; see photographs in Appendix C of Appendix F. Given the uncertainty 

surrounding edema as a relevant endpoint, the data could be analyzed both excluding and 

including edema as a response category for deformities. I doubt that the end result in terms of 

EC10 would differ substantially whether edema was included or not, but this would be worth 

evaluating.

In paragraph 2, page 7, (“yolk fry never swam up”), I am equally confused on this point and 

spent considerable time trying to make sense of it. The group LSV-2C had an original sample 

size of n=14 fish. Did the report authors remove the data from n=5 of these fish, where no 

hatched fry successfully swam up, in Figure 37 of Appendix F? (These data are also shown in 

Figure 3-1 on page 3-8 of Appendix A of Appendix F). This needs to be clarified in the report. It 

is important to note here that these five groups of embryos that did not swim up represent very

toxicologically meaningful data, since the toxicity of maternally transferred Se occurs mainly 

during yolk resorption. Were these five groups of embryos used for deformity frequency 

assessments? If not they should be. This was not clear to me from the report.

Final paragraph of section 5.0 (species sensitivity distribution): I agree with the reviewer’s 

comments and suggested reworking of Figure 6-1. As I am not familiar with the current USEPA 

policy on using EC20 vs. EC10 values, nor the use of the 5th percentile values as protective, I 

cannot make comment on these points. Obviously this is a very important point since it would 

appear to reduce the proposed site-specific Se criterion for the mine. 

A few comments on Appendix F of the report:

Appendix F, page 9 top: “As these non-selenium stressors for wild fish can also affect the test 

endpoints...”. This may be true, but without providing references to support this important 

claim it is meaningless. In the next paragraph “larval fish survival, growth, and deformities when 

no selenium exposure has occurred”. There is always Se exposure, as it is an essential ultra 

trace element. No Se exposure would have dire consequences. In the case of Se, it is the dose 

that defines the poison in both directions from “normal”!

Throughout the report, e.g. Figure 10 of Appendix F: “Egg mortality” should be changed to 

“Embryo mortality”. Eggs are gametes, they cannot die. Once the sperm fertilizes the egg, it is 

an embryo. 

Appendix F, Figure 25: it was unclear to me how egg Se and whole body Se were determined in 

the same fish to produce this relationship. I searched throughout the report and could find no 

methodological explanation. I assume that the mass of eggs removed for Se analysis was 

accounted for when measuring total Se in the remaining carcass?

6.0 Unaddressed Issues

Paragraph 1: I agree that there may be risks to aquatic bird species feeding in this area based 

on the comments of the reviewer and on the invertebrate Se concentrations shown in Table 13, 
7

Appendix A. Whether or not this has bearing on establishing a site-specific chronic Se criterion 

for the protection of fish I cannot say, as I am a biologist, not a regulator.

Paragraph 2-3: I somewhat disagree with these comments. Even if a comprehensive dataset of 

egg:whole body Se ratios were determined for the ten fish species inhabiting this area, we 

would still not know the relative sensitivities of those species to maternal Se exposure. I may be 

missing something here, but to me resources would be better spent conducting similar larval 

toxicity experiments in other fish species representing different taxa, e.g. cyprinids.

7.0 Recommended Recalculations

(I) Paragraph 1: I agree with the reviewer’s reassessment of the primary data and applaud 

her/him for the fine attention to detail with respect to “mining” out these discrepancies in the 

calculations. 

Paragraph 2: I strongly agree with this assessment, and as I stated above, consider the results 

obtained from these five batches of embryos to be biologically meaningful and essential to be 

included properly in the data analysis.

Paragraph 3: I agree that the results from these 10 samples should be excluded from the doseresponse analysis.

Paragraph 5 (top of page 12): From the outset I must state that I am not familiar with TRAP and 

its use in dose-response modeling, including the USEPA recommendations regarding its use. 

Thus I will assume the reviewer’s comments re: recommended curve fitting models for 

continuous vs. quantal data are true. Nonetheless, after viewing Figure 1 (and before reading 

the next paragraph on page 13) I observed that are no data between 100% and 0% survival, 

making this relationship tenuous at best. After subsequently reading the first paragraph on 

page 13, I agree with the reviewer’s assessment that a site-specific criterion based on this 

endpoint appears to fall “somewhere” between 14 and 27 μg/g dw, and that assigning a 

proposed criterion at a value greater than 14 μg/g would not be scientifically sound based on 

this endpoint.

(II) Overall, I agree with the assessment made by the reviewer for this first point. Table 2 in the 

document provides good evidence that, at least for larvae exposed to higher Se levels, not 

including deformed dead larvae will reduce the actual percentages of deformed larvae. 

However I agree with the report when they state that many of these larvae that died prior to 

swim-up would be difficult to assess. From my experience conducting similar research, this 

would be due to rapid fungal growth, decomposition and that dead larvae were likely also being 

fed on by live larvae. The authors of the report state that they sent approximately 100 “extra” 

larvae (from each of the high exposure fish LSV-003, 004, 005, 010 and 021) that died prior to 

swim-up to the contract lab that performed the deformity assessments. I agree with the 

reviewer that (a) this potential bias would not likely be as great for the fish with lower Se 

exposures, (b) that the bias may not greatly influence EC10/EC20 values, but (c) that the bias 

can only be in one direction: derivation of a less conservative criterion based on this endpoint.
8

For the second point in this section (starting middle of page 14), as I stated earlier in my review 

(page 5, above), I strongly agree with the reviewer’s comments. From what I can see in the 

report, I agree with the reviewer that each deformity category was assessed independently. In 

my opinion, which is in agreement with the reviewer, calculations should be based on the 

simple frequencies of fully normal vs. larvae exhibiting at least one deformity (i.e., not normal). 

For the data analysis made using the 2nd batch of hatchery trout (Table 3 and Figure 2), I am not 

in agreement with using those trout due to uncertainties related to husbandry, water 

temperature, nutrition, and other factors (possibly genetics as well) that differ between 

hatchery and wild environments. A proper field control is essential as stated earlier. With 

respect to the curve-fitting options and resulting output (described bottom of page 15), I will 

state again that I am not familiar with the use of TRAP and the appropriate model to use for 

such analysis. I would hazard a guess that use of either Least Squares Nonlinear Logistic 

Regression vs. Tolerance Distribution Analysis as curve-fitting models for this dataset would not 

appreciably change the resulting estimates of EC10 (6.5 μg/g egg dw) and EC20 (9.0 μg/g egg 

dw) values (i.e. both models would likely produce ECx values within 95% confidence intervals of 

each other).

The subsequent analysis (Figure 3), which excluded the hatchery trout but included the data 

from upstream sites (consisting of trout with elevated Se exposure), produced EC10 and EC20 

values of 15.2 and 17.5 μg/g egg dw, respectively. In the report (which also excluded hatchery 

trout and included upstream trout), EC10 and EC20 values were calculated as 17.7 and 21.6 

μg/g egg dw, respectively. Thus, estimated EC10 values from the report and FWS review vary by 

less than three-fold, ranging from 6.5 to 17.7 μg/g egg dw. Whether the EC10 of 6.5 μg/g egg 

dw is overly conservative or whether the EC10 of 17.7 μg/g egg dw is not protective for the 

sustainability of fish populations in this aquatic ecosystem cannot be ascertained until the 

dataset is expanded to include deformity frequencies from trout inhabiting a proper, 

regionally appropriate reference site (or sites). If this is undertaken, it is recommended that 

additional trout are collected from near field sites with “moderate” to “high” Se exposures to 

generate additional data for the “lower end” of the dose-response relationship for deformity 

frequency (i.e., between 10 to 40% normal fry).

Worth mentioning here is that a recently published article (DeForest et al. 2011) compiled 

available data for ten coldwater fish species that inhabit Canadian freshwater systems with the 

goal of creating a SSD for egg/ovary selenium. As suggested by the FWS reviewer in the 

document, this paper used mainly EC10 values for larval deformities and/or mortality, and the 

SSD was defined as the 5th percentile. The proposed guideline (Canadian regulatory term for 

criterion) using this approach was 20 μg/g egg dw. It should be noted that the EC10 values for 

brown trout and Yellowstone cutthroat trout that were derived in the present report were 

included as two of the ten species in DeForest et al. (2011). It also appears that the EC10 value 

for YCT in the report (35 μg/g egg) differs from the EC10 value shown in DeForest et al. (2011)

(25 μg/g egg).
9

(III) Determining species-specific relationships between egg/ovary and whole body Se 

concentrations are important, since previous studies (e.g., Holm et al. 2005) and the regionally 

relevant and comprehensive report included in my review package (Osmundson and Skorupa 

2011) document that there are significant differences among fish species in Se partitioning. This 

is not surprising given the diversity in life histories of female fish with respect to the pattern of 

oogenesis (e.g., synchronous vs. asynchronous [batch] spawners), length of the vitellogenic 

phase of oogenesis, amount of yolk in eggs, and even the large differences in amino acid 

sequence of vitellogenin and its catabolic products among fish species (specifically the number 

of methionine residues normally present in vitellogenin, which will be substituted for 

selenomethionine randomly and dose-dependently with increasing Se body burden). There is 

therefore much uncertainty involved in using generic equations (e.g., those provided by the 

USEPA in their 2004 draft Se criterion document) to predict whole body Se from egg/ovary Se in 

a given species.

In this section, when comparing Figure 4 of the FWS review to Figure 5-33 of the report it 

appears that the x and y axes were switched so that whole body Se is the dependent variable, 

and that only the brown trout Se relationship is shown. The R2 value for the brown trout 

regression changed from 0.81 in the report to 0.79 in the FWS review. I don’t have an opinion 

as to which way is correct, except to say that if whole body Se is to be predicted from egg Se, 

then the FWS review regression is more appropriate. 

In the review the p value for the regression is given as “p<0.0000001”, which seems unusual. 

Once p is less than 0.001 it indicates a highly significant result, and I’m not sure how much 

more confident one can be that the slope of this line is not horizontal with these extra zeroes

included.

Rearrangement of the regression axes resulted in a slightly lower predicted critical whole body 

Se concentration compared to the report (12.32 vs. 13.35 μg/g egg dw), based on the EC20 

value of 21.6 μg/g egg dw calculated in the report. Based on the EC10 value determined from 

Figure 2 in the FWS review (6.5 μg/g egg dw), the estimated critical whole body Se 

concentration was 5.5 μg/g egg dw. In my opinion, this whole body value appears overly 

conservative based on the literature and weight of evidence (see my comments in Summary

below). If this is to be included in the FWS review, it is suggested that back-calculation of the 

whole body values from the EC10 derived from Figure 3 of the FWS document is also included 

in this section, to provide some balance.

(IV) The first sentence of this section (“Tissue partitioning ratios...”) requires literature 

reference(s), since it is central to the argument presented. What evidence is there that Se 

partitioning to specific tissues such as the ovary changes with increasing exposure? I’m not 

disagreeing with this statement, but feel that it should be substantiated with evidence from the 

literature.

In this section the FWS reviewer makes a strong case that the use of brown trout Se partitioning 

among tissues to predict egg/ovary to whole body Se ratios for the other fish species making up 
10

the Crow Creek fish community may not be valid. Obviously there might even be more 

exceedances if a more conservative whole body criterion was used. It appears that sampling of 

the other fish species in this system, with collections occurring during mid to late vitellogenesis, 

to derive species-specific egg/ovary to whole body Se ratios would be appropriate from a 

regulatory standpoint.

Final Comments

Overall, I feel that the FWS reviewer provided accurate and fair comments based on sound 

scientific principles and statistical methodologies.

Although certain of the recommended recalculations produced only slight changes to derived 

values, which may seem semantic, they are based on sound science and should be considered 

by the authors of the report.

It is of utmost importance for an appropriate reference site(s), under no influence of 

anthropogenic activities that alter regionally normal aquatic selenium levels, be employed to 

obtain proper background frequencies of larval deformities in brown trout. I’ve done similar 

studies and had to go do an extra field season. It’s probably worth it.

The differences in opinion between the FWS reviewer and authors of the report represent 

common industry vs. government differences in conservatism with respect to conservation 

issues. What is appropriate criterion? In my opinion it is likely somewhere between the 6.5 μg 

Se/g egg dw EC10 value proposed by the FWS reviewer and the 17.7 μg Se/g egg dw EC10 value 

proposed in the report. Looking on the bright side, we are only talking a less than three-fold 

difference here; comparing this discrepancy to differences in protective criteria for the vast 

majority of inorganic and organic environmental contaminants in both ecological and human 

health risk assessments, it seems to me that a solution can be reached fairly easily with some 

additional resources devoted to this dataset. My goal in providing this review was hopefully to 

provide some balance to the argument from both perspectives (as a conservationist and a 

researcher who works in collaboration with industry).

Literature Cited

DeForest, D.K., Gilron, G., Armstrong, S.A. and Robertson, E.L. 2011. Species sensitivity 

distribution evaluation for selenium in fish eggs: Considerations for development of a 

Canadian tissue-based guideline. Integr. Environ. Assess. Manag. 7, 1-7.

Holm, J., Palace, V.P., Siwik, P., Sterling, G., Evans, R., Baron, C., Werner, J. and Wautier, K. 2005. 

Developmental effects of bioaccumulated selenium in eggs and larvae of two salmonid 

species. Environ. Toxicol. Chem. 24, 2373-2381.

Mueller, A.S., Klomann, S.D., Wolf, N.M., Schneider, S., Schmidt, R., Spielmann, J., Stangl, G., 

Eder, K. and Pallauf, J. 2008. Redox regulation of protein tyrosine phosphatase 1B by 

manipulation of dietary selenium affects the triglyceride concentration in rat liver. J. 

Nutr. 138, 2328-2336.
11

Palace, V.P., Halden, N.M., Yang, P., Evans, R.E. and Sterling, G.L. 2007. Determining residence 

patterns of rainbow trout using laser ablation inductively coupled plasma mass 

spectrometry (LA-ICP-MS) analysis of selenium in otoliths. Environ. Sci. Technol. 41, 3679–

3683.

Thomas, J.K. and Janz, D.M. 2011. Dietary selenomethionine exposure in adult zebrafish alters 

swimming performance, energetics and the physiological stress response. Aquat. Toxicol.

102, 79-86.

Wiseman, S., Thomas, J.K., McPhee, L., Hursky, O., Raine, J.C., Pietrock, M., Giesy, J.P., Hecker, 

M. and Janz, D.M. 2011. Attenuation of the cortisol response to stress in female rainbow 

trout chronically exposed to dietary selenomethionine. Aquat. Toxicol. 105, 643-651.
Dan Hoffman 

Geomorphologist/Hydrologist 

ATKINS 

1120 Cedar Street 

Missoula, Montana 

U.S.A. 

59802 

 November 6, 2011 

Mr Hoffman, 

I have completed my review and analysis of “Selected Technical Review 

Comments” by the US Fish and Wildlife Service (hereafter referred to as the 

FWS Commentary) relevant to a report entitled “Interpretive Findings for Field 

and Laboratory Studies and Literature Review in Support of a Site-Specific 

Selenium Criterion, Smoky Canyon Mine” (August 2010) (hereafter referred to as 

the Interpretive Report ) prepared by Formation Environmental and HabiTech 

for the J.R. Simplot Company in Pocatello, Idaho. In addition to reviewing the 

material contained within those documents, Appendices A,C and E in the 

Interpretive Findings Report, my review also included material presented in a 

report from the FY11 Environmental Contaminants Program Off-Refuge 

Investigations Sub-Activity (Project FFS ID: 6F50) entitled “CO-Selenium in Fish 

Tissue: Prediction Equations for Conversion between Whole Body, Muscle, and 

Eggs” (Osmundson and Skorupa 2011). I also reviewed several citations from 

the primary scientific literature to verify the accuracy of statements made in the 

FWS Commentary and the Interpretive Report.

In my capacity as a research scientist with the Department of Fisheries and 

Oceans, Canada I have been examining the potential impacts of selenium on 

freshwater fish for more than 12 years. I have a working knowledge of the 

statistics of dose-response modeling, and biology of salmonid fishes that are 

relevant to this area of study. While I bring this experience and working 

knowledge of the scientific literature, I must be clear that this review was 

performed outside of my responsibilities and working hours and, therefore, 

reflects my personal opinions and not necessarily those of the Department of 

Fisheries and Oceans. 

I have focused my review to address the following statement of work outlined in 

your message of October 4, 2011: 

“The reviewers will consider, but not be limited to, whether the authors’ conclusions

presented in the FWS technical comment document accurately and fairly reflect the

content of the reviewed documents and are scientifically valid and fair criticisms of the

reviewed documents.  If any of the criticisms presented in the FWS technical comment

document appear to be based on misunderstanding or unfair rendering of the reviewed

documents or are so unclear as to be undecipherable they should be identified and

included in the reviewers comments on the FWS technical comment document.  In
2

addition the reviewers should consider:  Are the specific critiques presented in the FWS

technical comment document coherently presented, logically sound, technically valid,

and nontrivial?  In addition, the reviewers should consider:  How could the FWS technical

comment document be improved?

The points of contention provided in the FWS Commentary are discussed in 

order of their appearance in that document. 

i) FWS Commentary

Section 1.0 Introduction – The FWS Commentary notes that the Yellowstone 

Cutthroat Trout (YCT) studies are portrayed as preliminary in the Introduction of 

the Interpretive Report and therefore, that interpretation is premature. However, it 

appears that analysis of the data from the YCT Adult Reproduction in the 

Interpretive Report are more than preliminary. Specifically, relationships between 

Se in YCT eggs and whole body Se, survival, growth and deformities appear to 

have been evaluated sufficiently to provide model analysis and Effects 

Concentration levels. I am unclear what analysis remain pending, such that the 

Introduction needs to portray the results as “preliminary”. 

I concur with the FWS Commentary that the Interpretive Report has 

misinterpreted the USEPA’s process for establishing a national water criterion 

(page 5, Interpretive Report). While only an aqueous exposure metric is 

nationally prescribed, this value was established with consideration that dietary 

exposure also occurred at the modeled Belews Lake field setting. Furthermore, 

the FWS Commentary correctly notes that USEPA’s proposed whole body tissue 

criterion(http://water.epa.gov/scitech/swguidance/standards/criteria/aqlife/polluta

nts/selenium/upload/complete.pdf) was established primarily using data from a 

laboratory study in which bluegill were exposed to dietary selenomethionine 

(Lemly, A.D. 1993. Metabolic stress during winter increases the toxicity of 

selenium to fish. Aquat. Toxicol., 27:133-158.) and not from a sole analysis of 

Belews Lake fish data. 

As the FWS Commentary contends, the 7.91 mg/kg dw whole body draft criterion 

established by USEPA was based on responses in bluegill exposed to 

temperature stress in addition to Se exposure. Therefore, comparison of genus 

mean chronic values (GMCV) from studies in which GMCV can be derived for 

trout, but for which no temperature stress is also included, is inappropriate. 

Specifically, the sensitivity of trout cannot be said to be less than bluegill based 

on the existing data in USEPA’s 2004 Draft Criterion. 

The Interpretive Report identifies that the USEPA commissioned the study by 

McIntyre et al. (2008) to re-examine Lemly’s winter stress syndrome paper of 

1993, but as noted in the FWS Commentary, there are important differences in 

the two studies that preclude their direct comparison. Among these are 

potentially different strains of fish (wild caught fish were used in Lemly (1993) 

while McIntyre used fish from a commercial source (Osage Catfisheries, Osage 
3

Beach, MO), differences in acclimation periods and temperature reduction, and 

different light regimes. The latter is noted in the FWS Commnetary and is 

especially important, because light cues are important for reproductive timing and 

energy metabolism shifts. The FWS Commentary also notes that lipid depletion 

determined in Lemly (1993) was not reported by McIntyre et al. (2008). Whether 

or not this was due entirely to different light regimes is uncertain, but direct 

comparisons of effects concentrations from the two studies certainly must 

consider these differences. 

Section 3.0 Literature Review 

Notwithstanding the dismissal of Besser et al. (2008) noted by in the FWS 

Commentary, other recent studies in which Se toxicity thresholds were 

developed were also omitted from the Interpretive Report. These omissions are 

important, because while the FWS Commentary contends that the “derivation of 

a proposed site-specific selenium criterion is not based substantively on the 

literature review”, the site specific thresholds that are developed in the 

Interpretive Report are considered in the context of literature values later in the 

document (eg. Section 6.2, Pages 68-69). 

Further to the reliance on Kennedy et al. (2000) for providing a toxicity threshold, 

it should be noted that while those authors reported a NOEC of > 81 mg SE/kg 

dw in eggs, two later studies using the same species reported effects 

thresholds of 16-20 mg SE/kg dw (Rudoph et al. 2008; Nautilus 2009). These 

values are more in line with the rest of the salmonid literature and do not suffer 

from some of the methodological issues we identified with Kennedy et al. in 

Palace and Hamilton (2001). I am in agreement with the statement in the 

Interpretive report, that egg survival is a highly variable endpoint and not 

particularly useful as a study endpoint for Se effects studies. 

4.0 Field Monitoring 

It is my understanding that there is a well established negative effect of sulphate 

on selenium uptake by aquatic macrophytes and lower trophic organisms and 

that competition for transport sites on the permease enzyme system is the 

accepted mechanism for this effect. However, a brief review of the literature 

indicates that, indeed, as stated in the FWS Commentary, there are few field 

validations of these laboratory results that would allow adjustment of a 

waterborne Se criterion based on elevated sulphate levels at field locations. 

The contention in the FWS Commentary that South Fork Tincup external 

reference site fish are significantly different from both upstream Crow Creek and 

Deer Creek reference sites is accurate. Specifically, Table 5.2 of the Interpretive 

Report shows that whole body Se in cutthroat trout from South Fork Tincup was 

2.56 mg/kg dw, while concentrations in the upstream reference sites range from 

ca. 8 to 25 mg/kg dw. References to support comparable Se concentrations in 

fish from other reference sites are provided in the FWS Commentary. In brown 
4

trout, reference fish from the Saratoga Hatchery contained whole body Se 

concentrations ranging from ca. 2.5 to 4.5 mg/kg dw, while the Crow Creek fish 

contained from 4.7 to 9.2 mg/kg dw (Tabel 5-2. Interpretive Report). What these 

comparisons of reference fish point out, is that fish from the “in field” reference 

sites (Upstream Crow Creek and Deer Creek) have higher than expected 

concentrations of Se. As contended in the FWS Commentary, this is not 

necessarily a surprising result. Fish in open systems are free to move and we 

have also shown evidence of salmonid fish movement into, and out of, Se 

impacted streams from reference locations (Palace et al. 2007, Environ. Sci. 

Technol. 41: 3679-3683). That peripyton more closely reflect Se exposure 

conditions at the chosen reference sites (Upstream Crow Creek and Deer 

Creeks) than reference fish captured from these locations, emphasizes the 

concept that fish movement is the likely source of higher than expected Se in fish 

at the reference sites locations. 

As the FWS Commentary contends, the use of fish densities in potentially Seaffected streams is not an appropriate assessment endpoint for the effects of Se 

(summary statement on Page 40 of the Interpretive Report). One reason for this 

is the potential for fish movement. In other words, streams with elevated Se can 

still contain fish assemblages even though effects on recruitment are likely to be 

present. Extending on the example from our research in Alberta that was cited 

above (Palace et al. 2007), spawning rainbow trout were still identified in Seimpacted streams, even though significant reproductive effects were identified 

(Holm et al. 2005, Environ. Toxicol. Chem. 24:2373-2381). The interpretation is 

that while recruitment may be affected, adult fish that are not adversely affected 

by elevated Se, continue to use the affected habitat for foraging and spawning 

(citation is, again, Palace et al. 2007). 

The FWS Commentary contends that demographic parameters such as those 

described in Vasterling (2003) would provide more appropriate metrics to identify 

potential effects of Se in affected streams. While I agree in principle, even these 

measures may provide responses that respond too slowly to provide meaningful 

“alarm bells” before population level impacts are evident. Continuing with our 

Alberta example above, population monitoring based on fish densities and size 

classes took more than 2 decades to show declining rainbow trout populations in 

the Se-affected Luscar Creek (G. Sterling, Alberta Sustainable Development, 

person. Commun. and VP Palace, GS Sterling, P Siwik, RE Evans, N Halden, 

KG Wautier, J Holm. 2006. Limitations of non-lethal sampling for determining 

spatiotemporal exposure to selenium in fish from mine impacted sites. 27th

Annual Society of Toxicology and Chemistry Meeting, Montreal, QC. Nov. 5-9). 

The timeframe for response variables in this type of demographic assessment is 

relevant when one considers the time to reproductive maturity (generally 3-5 

years) and long duration of active reproduction among cutthroat and brown trout 

(13+ years, Avery et al. (1985) [Eds] Sexual maturity and fecundity of brown trout 

in central and northern Wisconsin streams. Tech. bull.. Wisconsin Dept. of 

Natural Resources, No. 154, 12pp.). In simple terms, failures of several years 
5

classes would have to have occurred before demographic sampling would yield 

significant results because adult populations have extended reproductive 

lifespans. This is exacerbated in open systems where recruitment from adjacent 

unaffected systems can occur. Therefore, the most sensitive measures of 

potential Se impacts remain reproductive bioassays of relevant species under 

controlled conditions (Janz et al. 2010, Selenium Toxicity to Aquatic Organisms, 

Chapter 6 IN, Ecological Assessment of Selenium in the Aquatic Environment. 

Chapman et al. [Eds]. CRC Press, Boca Raton, FL.). 

5.0 Laboratory Studies 

As noted in the FWS Commentary, the adjustment of hatchery control responses 

to background (Upstream Crow and Deer Creeks) is inappropriate given the 

apparently elevated Se concentrations in these reference fish. Establishing a 

valid regression relationship is dependent on the inclusion of fish with a relevant 

range of Se concentrations. Specifically, it is fish at the lower end of the 

exposures that are most important for accurately deriving no effects 

concentrations (NOEC). Again, the FWS Commentary correctly points out that 

without these data points at the lower end of the exposures, overestimates of 

effects levels are inevitable. In my opinion, it is preferable to include the hatchery 

data uncorrected to strengthen the predictive capability of the regressions. 

The use of survival among 15 day post-swim up fry is not expected to be a 

sensitive endpoint for Se effects in a laboratory setting where rearing conditions 

are optimized. In contrast, the sum total of normal fry is appropriate as long as it 

does not include fry that did not swim up as part of the group characterized as 

“normal”. As noted in the FWS Commentary, failure to swim up at all is 

tantamount to death among wild fry. The fact that failure to swim-up occurred in a 

high proportion of eggs that contained high concentrations of Se (>25mg/kg dw, 

Figure 13 of the Interpretive Report) is an important result, albeit limited as a 

diagnostic indicator. Furthermore,15 day feeding trials using the fry with the 5 

highest concentrations of Se, and that failed to swim-up, is environmentally 

unrealistic and of no ecological value as noted in the FWS Commentary. 

As noted in my comment from the paragraph above, the sum total of normal fry is 

an appropriate diagnostic indicator for examining potential Se reproductive 

effects in fish. However, this is true only if records are maintained that allow the 

analyst to link deformity analyses for each fry across all 4 of the assessed 

categories and not simply a sum of the proportions of normal fry for each 

deformity category separately. The analysis description in the Interpretive 

Report, however, stipulates that: 

“deformities were evaluated as the sum fraction of normal fish (sum of 

normal fish/ total number of fish) for each deformity” (Appendix F, Page 

11). 

This indicates a de-coupling of the deformity analysis for each individual fry. As 

illustrated in the FWS Commentary, this practice allows for inflation of the total 
6

number of normal fry where multiple deformities do not exist. I view this point as 

being of utmost in importance for the derivation of reliable effects concentrations 

from the studies described in the Interpretive Report. In fact, as discussed below, 

I feel strongly enough about this shortcoming that I recommend a re-analysis of 

deformities from fry in these studies. 

I am fairly certain that re-plotting the regression of whole body Se and egg Se, 

with the X and Y axis reversed, would not significantly affect the derived whole 

body translations. 

The FWS Commentary contends that the sensitivity of brown trout is incorrectly 

illustrated in Figure 6-1 of the Interpretive Report because an EC20 is shown and 

not the preferred EC10 based on USEPA protocol. While this may be an 

important consideration, in my opinion it is more important to note that the data 

presented in the Interpretive Report are not comparable to the other maternal 

transfer studies. Again, this arises from the flawed deformity analysis that 

summed proportions of deformities independently. 

6.0 Unaddressed Issues 

I have limited expertise in the area of aquatic waterfowl biology and the potential 

for Se reproductive effects and, therefore, cannot comment on the first issue 

raised in the FWS Commentary. 

The Interpretive Report advocates the use of brown trout as an ecologically 

relevant species to monitor for potential Se effects. Summary statements in the 

Interpretive Report illustrate that it is the intent of the authors to use Se in whole 

body of brown trout to calculate potential Se concentrations in eggs of other 

species as suggested by the FWS Commentary. And so, it is important to 

address comments noted on page 9 of the FWS Commentary, regarding 

variability in the ratios of Se concentrations in whole body (or muscle) and those 

found in eggs among different fish species that have been investigated thus far. 

This is important when the relationships from one species are assumed to reflect 

the partitioning kinetics between tissues of other species within the same habitats 

and when that species is used as a sentinel species to determine protective 

effects levels. 
7

Egg Se (mg/kg dw)

Consider data in the figure above showing relationships between concentrations 

of Se in eggs and those in muscle for 8 freshwater species. (Source: Selenium 

Tissue Thresholds 2008. GEI Consultants, Golder Assoc., Parametrix, U of 

Sask.). 

The figure shows the regression plots for 8 species as well as 1:1 and 10:1 

regression lines. The Interpretive Report stipulates that for brown trout from this 

study, the ratio for egg to muscle Se is 1.54, meaning this species is near the 

bottom end of known regressions in terms of its slope. The FWS Commentary 

notes that this is also true when brown trout are considered among the 12 

documented species of fish in the Colorado River Basin for which these 

regressions have been documented. What this means is that, relative to most 

other species, brown trout accumulate higher concentrations of Se in muscle or 

whole body while maintaining lower Se concentrations in their eggs. Thus, if 

brown trout are used as the lone sentinel species for an aquatic system, an 

underestimate of potential reproductive effects from Se would result for some 

other fish species. 

7.0 Recommended Calculations 

(I.) Hatch to15-days post swim-up fry survival endpoint: 

I am uncertain why the endpoint “Hatch to 15 days post swim up fry survival” was 

calculated in the FWS Commentary by dividing %Total Survival by % Hatch. I 

have re-calculated the measure as follows, which seems to more closely agree 

with the data presented in the Interpretive Report: 

Hatch to 15 days post swim up survival = 100 - (%Hatch - %total survival) 

Using this equation to calculate the measure, there are still several 

miscalculations contained in Table 3-4 of Appendix A of the Interpretive Report. 

The following Table of revised values shows that 12 measures appear to be 

incorrect. These are highlighted in yellow and, in addition, there is one value 

(CC-150-016) for which total survival appears to be higher than the initial percent 

hatch (highlighted in red). While this is not possible (i.e. fish that did not hatch 

cannot survive to 15 days post hatch) the Interpretive Report provides a footnote 

explaining that “missing test organisms at test termination were not included in 
8

calculation” for the Hatch to 15 days post swim up survival. If this is the 

explanation for the total survival number being elevated above the initial hatch, 

this needs to be corrected before any regression analysis can be performed. 

TABLE OF RECALCULATED HATCH TO 15 DAY SURVIVAL VALUES

Sample ID % Hatch %Total Survival Hatch to 15 days survival 

SC-001 23.8 22.7 98.9

SC-002 22.8 22.3 99.5

SC-003 56.7 54.7 98

SC-004 31.1 27.3 96.2

SC-005 11.7 10.7 99

SC-006 93.2 92.7 99.5

SC-007 30.5 26.3 95.8

SC-008 66 65 99

SPC-001 99.3 98 98.7

SPC-002 100 100 100

SPC-003 97.5 94.2 96.7

SPC-004 100 100 100

SPC-005 98.8 96.8 98

SPC-006 99.7 96 96.3

CC-150-009 28.5 27 98.5

CC-150-011 96 95.3 99.3

CC-150-012 88.8 86 97.2

CC-150-013 66.7 57.5 90.8

CC-150-015 78.3 77.5 99.2

CC-150-016 14.8 15.4 100.6

CC-150-017 89.2 84.3 95.1

CC-150-018 87.2 84.3 97.1

CC-150-020 97.3 96.3 99

CC-350-006 71.7 67.7 96

CC-350-007 29.7 26 96.3

CC-350-008 67.7 62.6 94.9

LSV2C-002 99 96.6 97.6

LSV2C-003 93.5 2.25 8.75

LSV2C-004 50.6 16.8 66.2

LSV2C-005 71.3 23 51.7

LSV2C-006 0 - -

LSV2C-007 0 - -

LSV2C-008 88.3 84.2 95.9

LSV2C-010 87 11 24

LSV2C-012 98.3 95.7 97.4

LSV2C-016 95 91.7 96.7

LSV2C-017 71.3 60.5 89.2

LSV2C-019 94.2 88.9 94.7

LSV2C-020 89.2 86.8 97.6

LSV2C-021 69.3 14.8 45.5

The FWS Commentary is absolutely correct in noting that the 5 samples that did 

not produce swim-up fry at the end of the experimental period cannot be 

portrayed as anything other than 0% for legitimate regressions to be performed. 

Fry that fail to swim-up would be mortalities in the wild and portraying them in the 

same manner as fry which did swim-up from other clutches is completely 

inappropriate. 

The FWS notes several methodological errors and exclusions of data from 

missing organisms in the last paragraph on page 10. With regard to the hatchery 

control datasets, in my opinion, the exclusion of the first set of hatchery fish (SC) 

based on poor % hatch is appropriate. As noted in the Interpretive Report the 
9

poor performance of these eggs/fry may have been a result of different 

fertilization techniques or due to fungal infections. The second set of hatchery 

fish (SPC) is also problematic in terms of making comparisons to wild caught 

eggs because they were received as a later stage of development (eyed eggs) 

and not as fertilized eggs. Therefore, as the FWS Commentary points out, the 

percent hatch and potentially other parameters are not strictly comparable with 

other clutches of eggs treated differently. Finally, the FWS Commentary raises 

the issue of missing organisms from 10 groups of eggs (12 groups are actually 

noted as having missing organisms, but 2 of these are from the SC group which 

were excluded from further modelling). It is an unsubstantiated assumption that 

lost fry would have the same proportion of deformities and/or survival. 

Specifically, the Interpretive Report portrays these animals as having been lost 

“due to food clogging the drain pipe, resulting in an overflow of the test chamber.” 

In my opinion, it is more likely that dead or weakly swimming organisms would 

have been lost from the overflow at the top of the incubation container, because 

at this stage of development, fry are negatively phototropic and would seek 

refuge near the bottom. Therefore, as suggested by the FWS Commentary, it is 

more appropriate to exclude these tests from the developed models. This is also 

more consistent with the methodologies employed by the authors of the 

Interpretive Report in earlier stages of their studies. Specifically, they describe on 

page 3-6 of Appendix A that “Any organisms that were not found

during the test were considered dead”. This refers to the period of the test from 

fertilization to hatch, but the authors continue the statement to provide an 

alternative treatment of lost organisms in the second part of the studies (after 

swim-up) as: “....except during the 15-d swim-up study. For this phase of the 

study, any missing organisms were considered missing and were excluded from 

survival analysis.” 

With regard to the model chosen (Tolerance Distribution Analysis) that the FWS 

Commentary suggests for analyzing exposure-responses, I cannot provide input 

as I am unfamiliar with the USEPA’s TRAP software program and its conditions 

of operation and statistical assumptions, and requirements. I can, however, 

concur that the data generated for response variables from the brown trout study 

were not continuous, nor would they be expected to have infinite tails 

approaching 0 or 100%. In this sense, the description of the type of data 

generated by the studies is as described by the FWS Commentary on page 12, 

paragraph 1 (i.e. non-continuous with a finite threshold). 

(II.) Larval Deformity Endpoint: 

With regard to the collection of larval deformities, I concur with the FWS 

Commentary that exclusion of alevins and post swim-up fry that died before 

termination of the study is inappropriate and likely influenced deformity analysis 

by excluding potentially higher rates of deformed fry (dead fry are more likely to 

have been deformed) from further analysis. However, it should also be noted that 

many of the studies in the published literature have also excluded pre-swim up 

fry from analysis. In most cases, this exclusion is likely to have originated in an 
10

interest to maintain comparable developmental stages from each batch of eggs 

for assessment and comparison purposes. 

I have already discussed my views of the independent reporting of the 4 

categories of deformities for groups of fry rather than a summation of deformities 

for each individual fry (see the 3rd paragraph on Page 5 of this Report Under the 

heading 5.0 Laboratory Studies). It is worth noting that a recent publication 

(McDonald and Chapman 2009. Integ. Environ. Asses. Mngmt. 5:470–475) 

prescribes QA/QC procedures in which 

“A nonsequential labeling system is also recommended but will vary 

depending on the number of larval fish included in the deformity 

assessment. Individual vials are recommended for studies with small 

numbers of larval fish per adult female”. 

This type of re-analysis is, fortunately, possible where preserved specimens are 

available, as is the case for the studies described in the Interpretive Report. I 

would strongly favour such a re-analysis rather than the expression of 

“maximum fully normal fry” proposed by the authors of the FWS 

Commentary (page 14, 2nd paragraph). Should the authors of the Interpretive 

Report pursue such a re-analysis of the preserved specimens, they should 

consult the aforementioned publication for additional QA/QC measures (eg. Nonsequential labelling of fry, external deformity analysis checks) to ensure the most 

defensible data can be collected. 

Again, without familiarity regarding the operational requirements of the TRAP 

software, I am unable to comment on the model derivation offered by the authors 

of the FWS Commentary at the bottom of Page 15 and continuing on Page 16. 

However, as noted above, I am confident that the lack of a measure truly 

accounting for “fully normal fry” could certainly bias the EC10 and EC20 

estimates generated by the model toward the conservative. This is why, in my 

opinion, it is so critical to undertake a re-analysis of the preserved fry. Certain 

aspects of the studies described in the Interpretive Report cannot be reexamined (eg. Lost embryos during the 15 day swimup phase), but a reanalysis 

of the deformities is possible and should be the highest priority for this project. 

The issue of appropriate field control data has already been addressed on Page 

3 of this report, last paragraph). 

(III.) Regressing X on Y for Estimation of Critical Whole Body Tissue 

Concentrations 

As discussed earlier in my report there are significant species differences in 

distribution of Se between whole body or muscle and eggs. A review of 

Osmundson and Skorupa (2011) shows that deposition of Se into eggs from 

muscle or whole body can be expected to differ between wild fish and those 

raised in the laboratory. Similar data were also already discussed above in this 

report. While reasons for this difference are not fully know, it has been suggested 
11

that differential binding strengths of Se to the egg yolk precursor, vitellogenin, 

may account for some of the variability (Janz et al. 2010, Selenium Toxicity to 

Aquatic Organisms, Chapter 6 IN, Ecological Assessment of Selenium in the 

Aquatic Environment. Chapman et al. [Eds]. CRC Press, Boca Raton, FL.). 

Secondarily, there are marked differences in vitellogenin contents among 

different fish species’ eggs (Osmundson and Skorupa 2011). The regression plot 

and derivation of whole body Se from Se in eggs of brown trout that are given in 

the FWS Commentary (last paragraph of page 18 and first paragraph of page 19) 

are accurate. However, I contend, again, that a re-analysis of deformities in 

preserved fry and evaluation of the data from those analyses are far more 

preferable to expression of an EC10 based on the “fully normal fry” measure 

proposed in the FWS Commentary. Any derivation using these criteria are 

destined to be conservative in deriving an effect level, for reasons already 

discussed. 

 

(IV.) Test of Assumption that the 14 ppm Proposed Critical Whole Body 

Tissue Concentration Derived from Brown Trout Tissue Partitioning Data 

Would Generally Keep Egg Selenium Values Below a Critical Value of 21.63 

ppm Across a Multi-species Community of Fish:

 

Osmundson and Skorupa (2011) provide a data set illustrating significant 

differences among fish species in how they partition Se to their eggs/ovaries from 

muscle or whole body. The same report also shows differential partitioning to 

eggs/ovaries dependent on relative loads of Se among two species of fish (white 

sucker and green sunfish). For example, white sucker accumulate relatively low 

Se concentrations relative to those achieved by black bullhead. Moreovoer, even 

within white sucker, relatively less Se is deposited to their eggs from an 

increasing exposure dose during pre-spawning periods, relative to post spawning 

periods. These species and seasonal specific slopes for relating concentrations 

of Se in eggs to expected values in muscle or whole body become extremely 

important when whole body or muscle concentrations of Se are intended to 

provide monitoring data that will identify potential reproductive issues arising from 

Se exposure in the future. As identified in the FWS Commentary, the choice of 

brown trout as a (SSSC) monitoring species to derive a Site Specific criterion 

threshold value has great implications for risk management decisions for other 

species. Without a detailed accounting of complete fish assemblages in the study 

area described in the Interpetive Report, a judgement on the protectiveness of 

the brown trout SSSC is highly questionable. This is especially true given the 

lack of precision in deriving the brown trout EC10. 

Thank you for the opportunity to comment on both the Interpretive Report and the 

FWS Commentary. I am keenly interested in selenium and its potential to act as 

a reproductive toxicant and I would be pleased to offer additional opinions 

regarding re-analysis of deformities, should the stakeholders in this study deem 

that course of action to be appropriate. Should you have any questions regarding 
12

any of the opinions offered within this report, please do not hesitate to contact 

me. 

 Respectfully submitted, 

 

 Vince P. Palace, Ph.D. 

 

 

1 

Curriculum Vitae for Dr. Mark Rigby 

EDUCATION/EXPERIENCE

6/1/2008- Senior Risk Assessor, Parsons, South Jordan, USA. 

2001- Assistant Researcher at the University of California, Santa Barbara, CA, USA. 

6/2000-5/2008. Staff Scientist for Tetra Tech, Inc., Lafayette, CA, USA. 

2000. Lecturer (Parasitology) at the University of California, Santa Barbara, CA, 

USA. 

1999. Ph.D. Swiss Federal Institute of Technology, Zürich, Switzerland. 

1996. M.Sc. University of Alberta, Edmonton, Canada. 

1993. B.A. (distinction). University of California, Santa Barbara, USA. 

PUBLICATIONS IN REFEREED JOURNALS - ENVIRONMENTAL

1. Rigby MC, Deng X, Grieb TM, Teh SJ, Hung SSO. 2009. Effect threshold for 

selenium toxicity in juvenile Splittail, Pogonichthys macrolepidotus A. 

Bulletin of Environmental Contamination and Toxicology 84:76–79. 

2. Hartman B, Schumacher BA, Zimmerman JH, Springer DS, Elliot RJ, and 

Rigby MC. 2007. Results from EPA Funded Research Programs on the 

Importance of Purge Volume, Sample Volume, Sample Flow Rate and 

Temporal Variations on Soil Gas Concentrations. Proceedings of the AWMA 

Conference Vapor Intrusion: Learning from the Challenges (Sep 26-28, 2007; 

Providence, RI) 

3. Mills WB, Liu S, Rigby MC, Brenner D. 2007. Time-variable simulation of 

soil vapor intrusion into a building with a combined crawl space and 

basement. Environmental Science and Technology 41:4993-5001. 

PUBLICATIONS IN REFEREED JOURNALS - PARASITES

1. Costa G, Freitas N, Pinheiro de Carvalho MAA, Melo-Moreira E, Rigby MC. 

2009. Oncophora melanocephala (Nematoda, Camallanidae) from the chub 

mackerel, Scomber japonicus (Teleostei, Scombridae), caught off Madeira 

Island (Portugal). Acta Parasitologica 54:253-256. 

2. Vignon M, Sasal P, Rigby MC, Galzin R. 2009. Multiple parasite 

introduction and host management plan: case study of lutjanid fish in the 

Hawaiian Archipelago. Diseases of Aquatic Organisms 85:133–145. 

3. Rigby MC, Sharma RSK, Hechinger RF, Platt TR, Weaver JC. 2008. Two 

new species of Camallanus (Nematoda: Camallanidae) from freshwater turtles 

in Queensland, Australia. Journal of Parasitology 94:1364-1370. 

4. Moravec F, Justine J-L, Rigby MC. 2006. Some camallanid nematodes from 

marine perciform fishes off New Caledonia. Folia Parasitologica 53:223–239. 

5. Sharma R.S.K., Rigby M.C., Sumita S., Sani R.A., Vidyadaran M.K., and 

Jasni S. 2002. Redescription of Serpinema octorugatus (Nematoda: 

Camallanidae) from the Malayan box turtle Cuora amboinensis (Chelonia: 

Bataguridae). Systematic Parasitology 53: 19-28. 

6. Rigby M C, Hechinger, RF, Stevens L. 2002. Why should parasite resistance 

be costly? Trends in Parasitology 18:116-120. 

7. Rigby, M. C., and W. F. Font. 2001. Statistical re-analysis of the distinction 

between Spirocamallanus istiblenni and S. monotaxis (Nematoda: 

Camallanidae). Journal of Parasitology 87: 1210-1213. 
2 

8. Cribb TH, Pichelin S, Dufour V, Bray RA, Chauvet C, Faliex E, Galzin R, Lo 

CM, Lo-Yat A, Morand S, Rigby MC, Sasal P. 2000. Parasites of recruiting 

coral reef fish larvae in New Caledonia. International Journal for Parasitology 

30: 1445-1451. 

9. Font, W. F. and Rigby, M. C. 2000. Implications of a new Hawaiian host 

record from blue-lined snappers Lutjanus kasmira: is the nematode 

Spirocamallanus istiblenni native or introduced? Hawaii Biological Survey 

64: 53-56. 

10. Morand S, T. H. Cribb, M. Kulbicki, M. C. Rigby, C. Chauvet, V. Dufour, E. 

Faliex, R. Galzin, C.M. Lo, A. Lo-Yat, S. Pichelin, and P. Sasal. 2000. 

Endoparasite species richness of New Caledonian Butterfly fishes: host 

density and diet matter. Parasitology 121: 65-73. 

11. Jokela, J., Schmid-Hempel, P. and Rigby, M. C. 2000. Dr. Pangloss 

restrained by the Red Queen - steps towards an unified theory of defence. 

Oikos 89: 267-275. 

12. Rigby, M. C. and Jokela, J. 2000. Predator avoidance and immune defence: 

costs and trade-offs in snails. Proceedings of the Royal Society, Series B 

Biological Sciences 267: 171-176. 

13. Mark C. Rigby, Cédrik M. Lo, Thomas H. Cribb, Louis Euzet, Elisabeth 

Faliex, René Galzin, John C. Holmes, and Serge Morand. 1999. The role of 

parasites in coral reef fish communities – considerations and an example from 

French Polynesia with a parasite checklist. Cybium 23: 273-284. 

14. Morand, S., and Rigby, M. C. 1998. Cucullanin nematodes from coral reef 

fishes of French Polynesia, with a description of Cucullanus faliexae n. sp. 

(Nematoda: Chitwoodchabaudiidae). Journal of Parasitology 84: 1213-1217. 

15. Rigby, M. C., Adamson, M. L., and Deardorff, T. L. 1998. Camallanus 

carangis Olsen, 1954 (Nematoda: Camallanidae) reported from French 

Polynesia and Hawai’i with a redescription of the species. Journal of 

Parasitology 84: 158-162. 

16. Rigby, M. C., Holmes, J. C., Cribb, T. H., and Morand, S. 1997. Patterns of 

species diversity in the gastrointestinal helminths of a coral reef fish, 

Epinephelus merra (Serranidae), from French Polynesia and the South Pacific 

Ocean. Canadian Journal of Zoology 75: 1818-1827. 

17. Rigby, M. C., Font, W. F., and Deardorff, T. L. 1997. Redescription of 

Camallanus cotti Fujita, 1927 from Hawai’i. Journal of Parasitology 83: 

1161-1164. 

18. Rigby, M. C., and W. F. Font. 1997. Redescription and range extension of 

Spirocamallanus istiblenni Noble, 1966 (Nematoda: Camallanidae) from 

coral reef fishes in the Pacific. Journal of the Helminthological Society of 

Washington 64: 227-233. 

19. Rigby, M. C., and Adamson, M. L. 1997. Spirocamallanus species of French 

Polynesian coral reef fishes. Canadian Journal of Zoology 75: 1270-1279. 

20. Rigby, M. C., and V. Dufour. 1996. Parasites of coral reef fish recruits, 

Epinephelus merra (Serranidae), in French Polynesia. Journal of Parasitology 

82: 405-408. 

21. Rigby, M. C. 1996a. The epibionts of beach hoppers (Crustacea: Talitridae) 

of the North American Pacific coast. Journal of Natural History 30: 1329-

1336. 
3 

22. Rigby, M. C. 1996b. Association of a phoretic juvenile mite (Acari: 

Uropodidae) with beach hoppers (Crustacea: Talitridae). Journal of Natural 

History 30: 1617-1624. 

23. Adamson, M. L., and M. C. Rigby. 1996. Rhabditis (Crustorhabditis) 

stasileonovi (Belogurov) from beach hoppers (Amphipoda: Talitridae) from 

the Pacific coast of North America. Fundamental and Applied Nematology 19: 

579-584. 

NON-REFEREED PUBLICATIONS

1. Stephen J. Cornell, Yves Desdevises, and Mark C. Rigby. 1999. 

Evolutionary biology of host-parasite relationships: reality meets models. 

Trends in Ecology and Evolution 14: 423-425. 

BOOK CHAPTERS

1. Rigby, M. C., and Y. Moret. 2000. Life history trade-offs with immune 

defenses. In Evolutionary Biology of Host-Parasite Relationships, edited by R. 

Poulin, S. Morand and A. Skorping, eds., Amsterdam: Elsevier, pp 129-142. 

2. Rigby, M. C. 1999. The Camallanidae (Nematoda) of Indo-Pacific fishes: 

taxonomy, ecology, and host-parasite coevolution. In Proceedings of the 5th

Indo-Pacific Fish Conference (Nouméa, 3-8 November 1997), edited by B. 

Séret and J.Y. Sire. Paris: Société Française d’Ichtyologie & Institut de 

Recherche pour le Développement, pp 633-644. 

PRESENTATIONS – VAPOR INTRUSION

1. Hartman B., Schumacher B.A., Zimmerman J.H., Springer D.S., Elliot R.J., 

and Rigby M.C. 2007. Results from EPA Funded Research Programs on the 

Importance of Purge Volume, Sample Volume, Sample Flow Rate and 

Temporal Variations on Soil Gas Concentrations. AWMA Conference Vapor 

Intrusion: Learning from the Challenges (Sep 26-28, 2007; Providence, RI). 

2. Rigby, M.C., S. Liu, B. Mills, D. Brenner, S. Olliges, and T. H. Anderson. 

2005. New models for predicting indoor vapor concentrations for buildings 

with crawl spaces & basements: a case study from Moffet Field. Tri-Service 

Ecological Risk Assessment Work Group, January 24, 2006, Port Hueneme, 

CA. 

3. Rigby, M.C., B. Mills, S. Liu, K. Johnson, D. Brenner, S. Olliges, and T. H. 

Anderson. 2005. New models for predicting indoor vapor concentrations for 

buildings with crawl spaces & basements: a case study from Moffet Field. 

Society for Risk Analysis, December 5, 2005, Orlando, FL. 

4. Rigby, M.C., S. Liu, B. Mills, D. Brenner, S. Olliges, and T. H. Anderson. 

2005. Predicting indoor vapor concentrations for a building with a crawl 

space at Moffet Field. Groundwater Resources Association, May 25, 2005, 

San Jose, CA. 

5. C.B. Liban, K. Hancock, S. Liu, K. Johnson, M. Rigby. 2003. Calculating 

vapor intrusion to indoor air from groundwater and soil: an evolving science. 

PSAAPG and WRSPE, May 22, 2003, Long Beach, CA. 

6. K.M. Johnson and M.C. Rigby. 2003. Predicted and measured indoor vapour 

concentrations in housing over a VOC plume. Society for Risk Analysis, 

December 7-10, 2003, Baltimore, MD. 
4 

7. K.M. Johnson and M.C. Rigby. 2003. Predicted and measured indoor vapour 

concentrations in housing over a VOC plume at the Alameda Naval Air 

Station, California. Toxicology and Risk Assessment Conference, April 28 – 

May 1, 2003, Fairborn, OH. 

8. K.M. Johnson and M.C. Rigby. 2003. Predicted and measured indoor vapor 

concentrations in housing over a VOC plume at the Alameda Naval Air 

Station, California. Toxicology and Risk Assessment Conference, April 28 – 

May 1, 2003, Fairborn, OH. 

PRESENTATIONS – TOXICOLOGY

1. Torres, K.C., B.L. Kephart, D.F. Clark, M.F. Shibata, M.C. Rigby, T.E. Donn, 

and K.L. Waitman. 2002. Bioaccumulation of metals in terrestrial plants and 

invertebrates at Vandenberg Air Force Base, California. SETAC Annaul 

Meeting, November 2002, Salt Lake City, UT. 

2. D.F. Clark, B.L. Kephart, Torres, K.C., M.F. Shibata, M.C. Rigby, M. 

McElligott, and K.L. Waitman. 2003. Bioaccumulation of PCBs in terrestrial 

plants and invertebrates at Vandenberg Air Force Base, California. 

Toxicology and Risk Assessment Conference, April 28 – May 1, 2003, 

Fairborn, OH. 

3. K.M. Johnson and M.C. Rigby. 2004. Dicylcopentadiene toxicity 

assessment. Poster presentation at the Society for Risk Analysis, December 5-

8, 2004, Palm Springs, CA. 

PRESENTATIONS – PARASITES

1. Rigby, M. C. and Jokela, J. Costs of Predator avoidance and Environmental 

Stress in snails. Annual meeting SSE, ASN, SSB in Madison June 20-24, 

1999. 

2. M. C. Rigby and Y. Moret. Life history trade-offs and immune defense. 

Evolutionary biology of host-parasite relationship: reality meets models. 

Laboratoire Arago, Banyuls-sur-Mer, France, 3-7 May 1999. 

3. Rigby, M. C. Cantieni, M., and Jokela, J. Costs and interactions among 

defenses in snails. IIIrd cycle meeting: Life-history, immunocompetence and 

resistance against parasites. University of Neuchâtel, Neuchâtel, Switzerland, 

11-12 march 1999. 

4. Rigby, M. C. Cantieni, M., and Jokela, J. Costs and interactions among 

defenses in snails. Zoologia et Botanica, Swiss Society of Zoologists, Zürich, 

Switzerland, 17-19 February 1997. 

5. Rigby, M. C. and Jokela, J. Costs of Predator avoidance and Environmental 

Stress in snails. Annual meeting SSE, ASN, SSB in Vancouver June 20-24, 

1998. 

6. Rigby, M. C. Camallanid nematodes of Indo-Pacific coral reef fishes. 

Presented at the 5th Indo-Pacific Fish Conference, Nouméa, New Caledonia, 

3-8th November 1997. 

7. Font, W. F., and Rigby, M. C. Occurrence of Spirocamallanus istiblenni in 

tide pool and stream fishes of Hawai’i. Abstract of paper presented at the 

annual meeting of the Southeastern Society of Parasitologists, Greenville, 

South Carolina, April 17, 1997. 
5 

8. Rigby, M. C. and Holmes, J. C. Nematode parasites of fishes in the eastern 

south Pacific. Presented to the American Society of Parasitologists in June 

1996. (Title doesn’t say much, but it is about the possibility of multiple 

invasions of the nematode genus Spirocamallanus to French Polynesia.)
6 

REVIEWER FOR THE FOLLOWING JOURNALS/ORGANIZATIONS:

Folia Parasitologica 

The Journal of Parasitology 

Grant Agency of the Czech Republic 

Evolution 

Ecology Letters 

Marine Ecology Progress Series 

PROFESSIONAL AFFILIATIONS

American Society of Parasitologists (1995 – ) 

Helminthological Society of Washington (1996 -) 

American Institute of Biological Sciences (2000-) 

Society for Risk Analysis (2004-) 














































Vince Palace 

Employment: 

April 1999-Present Government of Canada, Department of Fisheries and Oceans 

Habitat Impacts Research Section

Position: Research Scientist 

 Director – Centre for Environmental Research on Pesticides

Responsibilities: Develop, design, and implement research projects to create and advance 

knowledge of factors affecting productivity and health of fish and fish 

habitat. Emphasis is on the biochemistry, histology and physiology 

measures of fish health and impacts on those by specific environmental 

contaminants, including pesticides. 

 Identify research opportunities, write research proposals, seek and foster 

collaborative research, and communicate research results in the primary 

scientific literature and at national and international conferences and 

meetings. 

 Supervise a research team and maintain large databases. 

 Train undergraduate and graduate students. 

 Critically review and evaluate manuscripts for scientific journals. 

 Represent the region and Department at national and international 

meetings.

 

September 1999- University of Manitoba, Departments of Zoology

Present And Environmental Studies

Position: Adjunct Professor and Lecturer 

Responsibilities: Coordinate, deliver and grade the 4th year Environmental Toxicology

course as well as Contribute to the Aquatic Biology Research Group

Graduate Student teaching 

Supervise graduate students 

Design, supervise and grade graduate level Topics in Zoology courses 

 
2

February 1999- Government of Canada, Department of Fisheries and Oceans 

March 1999 Habitat Management Division

Position: Impact Assessment Biologist, Alberta Area

Responsibilities: Reviewed projects potentially impacting freshwater fisheries resources 

under the framework of the Fisheries Act and the Canadian 

Environmental Assessment Act 

 Approved, provided advice for mitigating effects, or withheld 

authorization under the Fisheries Act 

 Conducted site inspections 

 Maintained public record of all relevant documents 

1995-1997 BEAR Environmental Science

Position: Consultant and Principal

Responsibilities: Acted as an independent consultant to review environmental impact 

statements for their thoroughness of biological effects monitoring 

considerations. 

1990-1994 Government of Canada, Department of Fisheries and Oceans 

Contaminants and Toxicology Research Section

Position: Project Leader, Great Lakes Lake Trout Study 

Projects: Research examined the mechanistic links between exposure to toxic chemicals, 

oxidative stress and reproductive failure in fish from the Great Lakes. High performance liquid 

chromatography (HPLC) techniques to quantify vitamins were developed. 

1991-1995 University of Manitoba, Department of Zoology 

Position: Teaching Assistant, Environmental Toxicology 

Responsibilities: Prepared lecture materials and evaluated students' oral and written course 

work. 
3

1987-1991 University of Manitoba, Department of Zoology 

Position: Teaching Assistant, Human Anatomy and Physiology 

Responsibilities: Lectured and demonstrated cellular, tissue level and whole body 

examinations of human anatomy and physiology; prepared and graded 

student examinations. 

Education:

July 1996- Manitoba Health Research Council Postdoctoral Fellow 

January 1999 Cardiovascular Research Institute 

St. Boniface Hospital, Winnipeg, Manitoba 

Project: The role of non-enzymatic antioxidants in the 

 pathogenesis and progression of heart failure. 

1992-1996 Doctor of Philosophy (Ph.D.) 

Department of Zoology 

University of Manitoba 

Thesis Title: Oxidative Stress in Lake Trout (Salvelinus namaycush) 

Exposed to Organochlorines that Induce the Phase I 

Biotransformation Enzyme System 

1988-1991 Masters of Science (MSc.)

Department of Zoology 

University of Manitoba 

Thesis Title: Superoxide Dismutase, Catalase and Glutathione Peroxidase 

Enzyme Activities as Indicators of Cadmium Exposure in 

Freshwater Fish from Small Canadian Pre-Cambrian Shield Lake 

Primary Current Research: 
4

Title: Effects based evaluation of current use pesticides on fish and fish 

habitat by the Center fro Environmental Research on Pesticides

Role: National Director 

Annual Budget: $1M/yr 

Project focus:

- carry out effects based research related to the potential impacts of pesticides 

on fish and fish habitat in marine and freshwater ecosystems.

- consult with the Pest management Regulatory agency (PMRA) and other 

federal departments to identify high priority research needs related to the 

environmental effects of pesticides on fish and fish habitat. 

- present results of effects based research to PMRA and other 

regulatory bodies in order to influence regulatory decisions in favour 

of increased protection for fish, fish habitat and marine ecosystems.

- provide advice to PMRA and other regulatory units on the potential 

for pesticides to impact fish and fish habitat. 

- Administer a research fund to support a network of pesticide expertise 

of regional research to address national pesticide priorities. 

Refereed Publications: 

Marvin CM, Tomy GT, Armitage JM, Arnot JA, McCarty L, Covaci A, Palace V. 2011. 

Hexabromocyclododecane: current understanding of chemistry, environmental fate and 

toxicology and implications for global management. Environ. Sci. Technol. 45, 8613–8623 

Tomy GT, Palace VP, Marvic C, Stapleton HM, Covaci A, Harrad S. 2011. Biotransformation 

of HBCD in Biological Systems Can Confound Temporal-Trend Studies. Environ. Sci. Tech. 45, 

364–365. 

Gemmill B, Pleskach K, Peters L, Palace V, Wautier K, Park B, Darling C, Rosenberg B, 

McCrindle R, Tomy GT. 2011. Toxicokinetics of tetrabromoethylcyclohexane (TBECH) in 

juvenile brown trout (Salmo trutta) and effects on plasma sex steroids. Aquat. Toxicol. 101:309-

317. 

Friedrich LA, Orr PL, Halden NM, Yang P, Palace VP. 2011. Exposure histories derived from 

selenium in otoliths of three cold-water fish species captured downstream from coal mining 

activity. Aquat. Toxicol. 105:492-496. 

Marteinseon SC, Kimmins S, Letcher RJ, Palace VP, Bird DM, Ritchie IJ, Fernie KJ. 2011. Diet 
5

exposure to technical hexabromocyclododecane (HBCD) affects testes and circulating 

testosterone and thyroxine levels in American kestrels. (Falco sparverius). Environ. Res. 

111:1116-1123. 

Park BJ, Palace V, Wautier K, Gemmill B, Tomy G. 2011. Thyroid axis disruption in juvenile 

Brown trout (Salmo trutta) exposed to the flame retardant -tetrabromocyclohexane (-TBECH) 

via the diet. Environ. Sci. Technol. 45:7923-7927. 

Boudreau HS, Rossiter JP, Williams LD, Palace VP, RossGM. 2009. The Effects of Metal Ion 

Exposure with -synuclein-like immunoreactivity in the Central Nervous System of Fish, 

Catostomus commersoni. Aquat. Toxicol. 92:258-263 

Zhang S, Bursian SJ, Martin PA, Chan HM, Tomy G, Palace VP, Mayne GJ, Martin JW. 2009. 

Reproductive and developmental toxicity of a pentabrominated diphenyl ether mixture, DE-71, 

to ranch mink (Mustela vison) and hazard assessment for wild mink in the Great Lakes region. 

Toxicol. Sci. 110:107-116. 

Johnson KE, Park BJ, Palace VP, Ollson CA, Reimer KJ. 2009. Biomarkers as indicators of 

marine ecosystem recovery at a polychlorinated biphenyl contaminate site at Saglek Bay, 

Labrador. (????) 

Miller L, Rasmussen JB, Palace VP, Hontela A. 2009. The physiological stress response and 

oxidative stress biomarkers in rainmbow trout and brook trout from selenium impacted streams 

in a coal mining region. J. Applied Toxicol. 29:681-688. 

Godard D, Park BJ and Palace VP. 2009. Histopathological assessment of the sub-lethal effects 

of instantaneous pressure changes (IPCs) on the internal organs of fish. Environ. Sci. Technol. 

(Submitted Nov. 2010) 

Godard D and VP Palace. 2008. A review of the evolution, structure, development and function 

of the teleost swimbladder and factors governing its susceptibility to pressure induced effects 

from underwater detonations. DFO Technical report (In preparation) 

Miller, LL, Rasmussen JB, Palace VP, Hontela A. 2009. The physiological stress response in 

white suckers from agricultural drain waters containing pesticides and selenium. Ecotox. 

Environ. Saf. 72:1249–1256 

Goodmanson, AT, VP Palace and D Gillis. 2008. Assessing site fidelity of golden shiners 

(Notemigonous chrysoleucas) ina riverine environment using mark-recpature analysis and Monte 

Carol data simulation. DFO Technical Report (In preparation). 

Palace, VP, BJ Park, K Pleskach, B Gemmill and GT Tomy. 2009. Biotransformation enzymes 

and thyroid axis disruption in juvenile Rainbow trout (Oncorhynchus mykiss) exposed to HBCD 

Diastereoisomers. Environ. Sci. Tech. 42:1967-1972. 
6

Palace VP, RE Evans, KG Wautier, KH Mills, PJ Blanchfield, BJ Park, CL Baron and KA Kidd. 

2008. Interspecies differences in biochemical, histopathological, and population responses in 

four wild fish species exposed to ethynyestradiol added to a whole lake. Can. J. Fish. Aquat. 

Sci. 66:1920-1935. 

Palace VP, K Pleskach, T Halldorson, R Danell, K Wauiter, RE Evans, M Alaee, C Marvin and 

GT Tomy. 2007. Biotransformation enzymes and thyroid axis disruption in juvenile rainbow 

trout (Oncorhynchus mykiss) exposed to hexabromocyclododecane diastereoisomers. Environ. 

Sci. Technol. 42:1967-1972. 

Werner J, V Palace, R Shiu and A Yarmil. 2008. A real-time PCR method for the quantification 

of the two isoforms of metallothionein in lake trout (Salvelinus namaycush). Arch. Environ. 

Contam. Toxicol. 54:84-91. 

Couillard CM, RW Macdonald, SC Courtenay and VP Palace. 2007. Interactions between toxic 

chemicals and other environmental factors affecting the risk of impacts on aquatic organisms: a 

review with a Canadian perspective – Interactions affecting exposure. Environmental Reviews. 

16:1-17. 

Palace VP, NM Halden, P yang, RE Evans, G sterling. 2007. Determining residence patterns of 

rainbow trout using laser ablation inductively coupled plasma mass spectrometry (LA-ICP-MS) 

analysis of selenium in otoliths. Environ. Sci. Tech. 41:3679-3683. 

Martin,P.A., Mayne, G.J., Bursian,S.J., Tomy, G., Palace, V., Pekarik, C., and Smits, J. 2007. 

Immunotoxicity of the commercial polybrominated diphenyl ether mixture DE-71 in ranched 

mink. Environ Toxicol Chem. 26: 988-997. 

Kidd KA, PJ Blanchfield, KH Mills, VP Palace, RE Evans, J Lazorchak and B Flick. 2007. 

Collapse of a fish population following exposure to a synthetic estrogen. PNAS. 104:8897-8901. 

Miller LL, F Wang, VP Palace and A Hontela. 2007. Effects of acute and subchronic exposures 

to waterborne selenium on the physiological stress response and oxidative stress indicators in 

juvenile rainbow trout. Aquat. Toxicol. 83:263-271. 

Fudge TS, KG Wautier and VP Palace. 2008. Escapement success of rainbow trout 

(Oncorhynchus mykiss) fry from artificial redds with different fine sediment loadings. NAJFM. 

28:758–765. 

Tomy GT, VP Palace, K Pleskach, N Ismail, T Oswald, R Danell, K Wautier, RE Evans. 2007. 

Dietary exposure of juvenile rainbow trout (Oncorhynchus mykiss) to 1,2(bis(2,4,6-

tribromophenoxy)ethane (BTBPE): bioaccumulation parameters and biochemical effects. 

Environ. Sci. Technol. 41:4913-4918. 

Palace VP, S Kollar, RE Evans, LE Peters, CL Baron, KG Wautier and M Parsosn. 2007. An 
7

initial assessment of arsenic mobilization in an area of historic gold mining in Nova Scotia, 

Canada and bioaccumulation of arsenic and mercury in resident American Eels (Anguilla 

rostrata). Geochem. Explor. Environ. Anal. (Invited paper, In Press). 

Law K, VP Palace, T Halldorson, R Danell, K Wautier, RE Evans, M Alaee, C Marvin, G 

Tomy. 2006. Dietary accumulation of hexabromocyclododecane diastereoisomers in juvenile 

rainbow trout (Oncorhynchus mykiss) I: bioaccumulation parameters and evidenc of 

bioisomerization. Environ. Toxicol. Chem. 25: 1757-1761. 

Palace VP, Werner J. 2006. Vitamins A and E in the Maternal Diet Influence Egg Quality and 

Early Life Stage Development in Fish: A Review. Proceedings of the 29th Larval Fish 

Conference, Barcelona Sp, July 11-14, 2005. Scientia Marina.70S2:41-57 

Klaverkamp JF, VP Palace, CL Baron, RE Evans, KG Wautier. 2006. Cumulative effects of 

multiple contaminants on caged fish. Can Wat. Qual. Res. J. 41:244-255. 

Doebel C, Baron CL, Evans RE, Palace VP. 2005. The use of small bodied forage fish in caged 

exposures for metal mining environmental effects monitoring. Proceedings of the International 

Conference on Heavy Meals in the Environment. Rio DeJaneiro, Brazil. June 5-9, 2005. (In 

press). 

Gray M, Munkittrick K, Dahl J, Baron C, Wautier K, Kollar S, Werner J, Palace V. 2005. 

METALS and metallothionein in slimy sculpin (Cottus cognatus) captured at a Diamond mining 

Site in Northern Canada. Proceedings of the International Conference on Heavy Meals in the 

Environment. Rio DeJaneiro, Brazil. June 5-9, 2005. (In press). 

Werner J, Wautier K, Mills K, Chalanchuk S, Kidd K, Palace VP. 2005. Reproductive fitness of 

lake trout (Salvelinus namaycush) exposed to environmentally relevant concentrations of the 

potent estrogen ethynylestradiol (EE2) in a whole lake exposure experiment. Proceedings of the 

29th Larval Fish Conference, Barcelona Sp, July 11-14, 2005 Scientia Marina, 70S2:59-66. 

Hill MF, Palace VP, Kaur K, Kumar D, Khaper N, Singal PK. 2005. Reduction in oxidative 

stress and modulation of heart failure subsequent to myocardial infarction in rats. Exper. Clin. 

Cardiol. 10:146-153. 

Palace VP, C Doebel, CL Baron, RE Evans, KG Wauiter, JF Klaverkamp, J Werner, S Kollar. 

2005. Caging small bodied fish as an alternative method for environmental effects monitoring 

(EEM). Can. Wat. Qual. Res. J. 40:328-333. 

Palace VP, Wautier KG, Evans RE, Blanchfield P, Mills K, Chalanchuk S, Godard D, McMaster

ME, Tetrault G, Peters LE, Vandenbyllaardt L and Kidd KA. 2006. Biochemical and Biochemical 

and Histopathological Effects of Ethynylestradiol in Pearl Dace (Margariscus 

margarita)Chronically Exposed to the Synthetic Estrogen in a Whole Lake Experiment. Environ. 
8

Toxicol. Chem.. 25:260-271. 

Holm J, VP Palace, P. Siwik, G Sterling, RE Evans, CL Baron, J Werner, K Wautier. 2005. 

Developmental effects of bioaccumualted selenium in eggs and larvae of two salmonid species. 

Environ. Toxicol. Chem. 24:2373-2381. 

Bauder MB, Palace VP, Hodson PV. 2005. Is oxidative stress the mechanism of blue sac disease 

in retene-exposed trout larvae? Environ. Toxicol. Chem. 24:694-702. 

Spallholz JE, Boylan LM, Palace VP, Chen J, Smith L, Rahman MM, Robertson JD. 2005. 

Arsenic and selenium in human hair: a comparison of five countries with and without 

arsenicosis. Biological Trace Element Research. 106(galley stage) 

Martin PA, Mayne G, Bursian S, Palace VP, Kurunthachalam, K. 2005. Thyroid and vitamin A 

status in mink fed PCB-contaminated carp from Saginaw Bay, Michigan. Environ. Res. 101:53-

67. 

Dwillow R, Palace V, Kaur K, Singal PK. 2004. Regional changes in metallothioneins during 

global ischemia-reperfusion in rats. In Heart Disease- pathogenesis, Diagnosis and Treatmetn. A. 

Kimchi [Ed]. Medimond International Proceedings, Bologna, Italy, pp. 327-333. 

Tomy GT, Palace VP, Halldorson T, Braekevelt E, Danell R, Wautier K, Evans B, Brinkworth 

L, Fisk AT. 2004. Bioaccumulation, biotransformation and biochemical effects of brominated 

diphenyl ethers in juvenile lake trout (Salvelinus namycush). Environ. Sci. Technol. 38:1496-

1504. 

Tomy,GT, SA Tittlemier, VP Palace, E Braekevelt, WR Budakowski, B Lau, G Eales, JC Plohman. 

2003. Biotransformation of Perfluoroooctane Sulfonamide and Ethyl Perfluoroooctane 

Sulfonamide by Rainbow Trout (Onchorhynchus mykiss) Liver Microsomes. Environ. Sci. 

Technol..38:758-762. 

Brown, SB, RE Evans, L Vandenbyllardt, KW Finnson, VP Palace, AS Kane, AY 

Yarechewski, DCG Muir. 2003. Altered Thyroid Status in Lake Trout (Salvelinus namaycush) 

Exposed to Co-planar 3,3',4,4',5-Pentachlorobiphenyl. Aquat. Toxicol. 67:75-85.

Doebel C, CL Baron, RE Evans, KG Wautier, J Werener and VP Palace. 2004. The potential for 

using caged pearl dace (Semotilus margarita) for metal mining Environmental Effects 

Montioring. (EEM). 2003. Bull. Environ. Contam. Toxicol. 72:409-414. 

Palace, VP. 2003. Vitamin E as a mediator and indicator of toxicant induced stress in fish. In 

Vitamin E in Health and Disease. K. Dabrowski [Ed]. (In Press:Invited Paper). 

Palace, VP, CL Baron, RE Evans and K Wautier. 2003. Mobility of Arsenic in a Canadian 

Freshwater System Receiving Gold Mine Effluents. J. Physique IV 107:1009-1012. 
9

Palace, VP, CL Baron, RE Evans, K Wautier and JF Klaverkamp. 2003. Metals, Metalloids and 

Metallothionein in Tissues of Fish from A Canadian Freshwater System Receiving Gold Mining 

Effluents and in Fish from Nearby Reference Areas. J. Physique IV 107:1005-1008. 

Palace, VP, JE Spallholz, J Holm, K Wautier, RE Evans and CL Baron. 2004. Metabolism of 

selenomethionine by rainbow trout (Oncorhynchus mykiss) embryos can generate oxidative 

stress. Ecotoxicol. Environ. Saf. 58:17-21 

Palace, VP, RE Evans, K Wautier, CL Baron, S Kollar, J Werner, J Holm, P Siwik, G Sterling, 

CF Johnson. 2003. An assessment of the potential for selenium to impair reproduction in bull 

trout (Salvelinus confluentus) from an area of active coal mining. Environ. Biol. Fishes. (In 

Press). 

Spallholz, JE, VP Palace, TW Reid. 2004. Methioninase and selenomethionine but not Semethylselenocysteine generate methyselenol and superoxide in an in vitro chemiluminescence 

assay: experiments and review. Biochem. Pharamacol. 67:547-554. 

Nicholas, D, S Ramamoorthy, VP Palace, S Spring, J Moore, RF Rosenzweig. 2002. 

Biogeochemical Transformations of Arsenic in Circumneutral Freshwater Sediments. 

Biodegredation 14(2):123-37. 

Holm, J, VP Palace, K Wautier, RE Evans, CL Baron, C Podemski, P Siwik, G Sterling. 2002. 

An assessment of the development and survival of wild rainbow trout (Oncorhynchus mykiss) 

and brook trout (Salvelinus fontinalis) exposed to elevated selenium in an area of active coal 

mining. In The Big Fish Bang. Proceedings of the 26th Annual Larval Fish Conference, July 22-

26, Bergen, Norway. 257-273.

Majewski, AR, PJ Blanchfield, VP Palace, K Wautier. 2002. Waterborne 17-ethynlestradiol 

affects aggressive behaviour of male fathead minnows (Pimephales promelas) under artificial 

spawning conditions. Wat. Qual. Res. J. Can. 37:697-710. 

Palace, VP, RE Evans, K Wautier, L Vandenbyllardt, W Vandersteen, K Kidd. 2002. Induction 

of vitellogenin and histological effects in wild fathead minnows from a lake experimentally 

treated with the synthetic estrogen, ethynylestradiol. Wat. Qual. Res. J. Can. 37:637-650. 

Werner, J, K Wautier, RE Evans, CL Baron, K Kidd, VP Palace. 2002. Waterborne 

ethynylestradiol induces vitellogenin and alters metallothionein in lake trout (Salvelinus 

namaycush). Aquat. Toxicol. 62:321-328. 

Danelisen I, Palace VP, Lou H, Singal PK. 2002. Maintenance of myocardial levels of vitamin 

A in heart failure due to adriamycin. J.Mol.Cell.Cardio. 34(7):789-95.

Palace, VP, RE Evans, K Wautier, CL Baron, J Werner, JF Klaverkamp, KA Kidd, TA Dick. 

2001. Altered Metabolism of Lipid Soluble Antioxidant Vitamins in Juvenile Sturgeon Exposed 

to Waterborne Ethynylestradiol. Environ. Toxicol. Chem. 20:2370-2376. 
10

Hamilton, S J, VP Palace. 2001. Assessment of Selenium Effects in Lotic Ecosystems. Ecotox. 

Environ. Saf. 50:161-166.

Palace VP, Wautier K, Evans RE, Baron CL, Werner J, Ranson CR, Klaverkamp JF Kidd 

K.2001. Effects of 17- Estradiol Exposure on Metallothionein and Fat Soluble Antioxidant 

Vitamins in Juvenile Lake Trout (Salvelinus namaycush). Bull. Environ. Contam. Toxicol. 

66:591-596.

Palace, VP, SM Allen-Gil, SB Brown, RE Evans, DA Metner, DH Landers, LR Curtis, WL 

Lockhart. 2001. Altered vitamin A metabolism in arctic grayling (Thymallus arcticus) exposed 

to 3,3',4,4'-tetrachlorobiphenyl. Chemosphere . 45:185-193.

Palace, VP, MF Hill, F Farahmand, PK Singal. 1999. Mobilization of antioxidant vitamin pools 

and hemodynamic function following myocardial infarction. Circulation. 99:121-126. 

Palace, VP, N Khaper, Q Qin, PK Singal. 1999. Antioxidant potentials of vitamin A and 

carotenoids and their relevance to heart disease. Free Rad. Biol. Med. 26:746-761. 

Palace, VP, D Kumar, MF Hill, N Khaper, PK Singal. 1999. Regional differences in nonenzymatic antioxidants in the heart under control and oxidative stress conditions. J. Mol. Cell. 

Cardiol., 31: 193-202, 1999. 

Pichardo, JP, VP Palace, F Farahmand, PK Singal. 1999. Myocardial oxidative stress changes 

during right heart failure in rats. Mol. Cell. Biochem, 196:51-57. 

Palace, VP, MH Hill, N Khaper, PK Singal. 1999. Metabolism of vitamin A in the heart 

increases after a myocardial infarction. Free Rad. Biol. Med. 26:1501-1507. 

Singal, PK, N Khaper, VP Palace, D Kumar. 1999. Oxidative stress in heart failure: current 

understanding and prospective. Heart Failure Reviews, 4(2): 1-10.

Singal, PK, N Khaper, VP Palace, D Kumar. 1999. On the role of coenzyme Q10 in 

cardiovascular diseases. Cardiovasc. Res. 43:250-251. 

 

Singal, PK, N Khaper, VP Palace, D Kumar. 1998. The role of oxidative stress in the genesis of 

heart disease. Cardiovasc. Res. 40:426-432.

Iliskovic, N, T Li, N Khaper, VP Palace, PK Singal. 1998. Modulation of adriamycin-induced 

changes in serum free fatty acids, albumin and cardiac oxidative stress. Mol. Cell. Biochem.

188:161-166. 

Singal, PK, RA-A Ghani, N Khaper, VP Palace, MF Hill. 1997. Antioxidant adaptations and 
11

cardiac dysfunction: involvement of vitamin E. J. Adapt. Med. 1:5-15.

Palace, VP. SB Brown, CL Baron, J Fitzsimmons, JF Klaverkamp. 1998. Relationships between 

induction of the mixed function oxidase enzyme system and oxidative stress: relevance for lake 

trout from Lake Ontario and early mortality syndrome (EMS) of their offspring. American 

Fisheries Society Symposium. 21:146-153.

Brown, SB, J Fitzsimons, VP Palace, L Vandenbyllardt. 1998. Thiamine and early mortality 

syndrome in lake trout. American Fisheries Society Symposium 21:18-25.

Palace, VP, CL Baron, JF Klaverkamp. 1998. An assessment of Ah-inducible Phase I and Phase 

II enzymatic activities and oxidative stress indices in adult lake trout (Salvelinus namaycush) 

from Lake Ontario and Lake Superior. Aquatic Toxicology 42:149-168. 

Palace, VP, SB Brown, CL Baron, J Fitzsimmons, B Woodin, JJ Stegeman, JF Klaverkamp. 

1998. An evaluation of the relationships among oxidative stress, antioxidant vitamins and early 

mortality syndrome (EMS) of lake trout (Salvelinus namaycush) from Lake Ontario. Aquat. 

Toxicol. 43:195-208. 

Palace, VP, JF Klaverkamp, CL Baron, SB Brown. 1997. Metabolism of 3

H-Retinol by lake 

trout (Salvelinus namaycush) pre-exposed to 3,3',4,4',5-pentachlorobiphenyl (PCB 126). Aquat. 

Toxicol. 39:321-332. 

Palace, VP, TA Dick, SB Brown, CL Baron, JF Klaverkamp. 1996. Oxidative stress in lake 

sturgeon (Acipenser fulvescens) orally exposed to 2,3,7,8-tetrachlordibenzofuran. Aquat. Toxicol.

35:79-92. 

Palace, VP, SB Brown, DA Metner, WL Lockhart, DCG. Muir, JF Klaverkamp. 1996. Mixed 

function oxidase enzyme activity and oxidative stress in lake trout (Salvelinus namaycush)

exposed to 3,3'4,4',5-pentachlorobiphenyl. Environ. Toxicol. Chem. 15(6):955-960. 

Palace, VP, SB Brown. 1994. HPLC determination of tocopherol, retinol, dehydroretinol and 

retinyl palmitate in tissues of lake char (Salvelinus namaycush) exposed to coplanar 3,3',4,4',5-

pentachlorobiphenyl. Environ. Toxicol. Chem. 13:473-476. 

Palace, VP, JF Klaverkamp. 1993. Variation of hepatic enzymes in three species of freshwater 

fish from precambrian shield lakes and the effect of cadmium exposure. Comp. Biochem. 

Physiol. 104C:147-154. 

Palace, VP, HS Majewski, JF Klaverkamp. 1993. Interactions among antioxidant defenses in 

liver of rainbow trout (Oncorhynchus mykiss) exposed to cadmium. Can. J. Fish. Aquat. Sci.

50:156-162. 

Palace, VP, HS Majewski, JF Klaverkamp. 1990. Effects of sampling and storage conditions on 
12

ther stability of biochemical parameters measured in rainbow trout (Oncorhynchus mykiss)

liver. Comp. Biochem. Physiol. 95B:51-55. 

Book Chapters and Reports: 

Palace VP. 2011. Investigating selenium toxicity using an environmental effects monitoring 

approach. Proceedings of the Investigation of casue Workshop, Gatineau QC. April 29-30, 2009. 

Palace VP, Baron CL, Bandler C. 2010. An update on the use of caged fish for EEM. 

Proceedings of the Environmental Effects Monitoring Investigation of Cause Workshop, 

Gatineau QC. April 29-30, 2009. 

David M Janz, Marjorie L Brooks, Peter M Chapman, David K DeForest, Guy Gilron, Dale 

Hoff, William A Hopkins, Dennis O McIntyre, Christopher A Mebane, Vincent P Palace, 

Joseph P Skorupa, Mark Wayland. 2009. Chapter 6: Selenium Toxicity to Aquatic Organisms. 

Society of Environmental Toxicology and Chemistry Pelston Workshop. SETAC Press, 

Pensacola FL. 

Godard, DR, L Peters, R Evans, K Wautier, PA Cott, B Hanna and VP Palace. 2008. 

Development of Histopathology Tools to Assess Instantaneous Pressure Change-Induced Effects 

in Rainbow Trout (Onchorhynchus mykiss) Early Life Stages. Department of Fisheries and 

Oceans Environmental Science Strategic Research Report. (In Press). 

Khaper, N, VP Palace, MF Hill, D Kumar, PK Singal. 1999. Heart failure: a failed adaptation to 

oxidative stress. In: Adaptation Biology and Medicine. Vol. 2. Narosa Publishers, New Delhi, 

India. Pp. 14-29. 

Kowaluk, B, N Khaper, C Rigatto, VP Palace, PK Singal. 1998. Mechanisms of cardioprotective 

effects of propanolol against reperfusion injury. In: Ischemic Heart. Eds. S. Mochizuki, N. 

Takeda, M. Nagano and N.S. Dhalla. Kluwer Academic Publishers, Boston. pp. 367-378. 

Invited Lectures 

VP Palace, CL Baron, RE Evans KG Wautier, C Doebel, S Kollar and L Brinkworth. 2006. Examination of the 

Relationship Between Primary Productivity and Arsenic Mobility In Balmer Lake. University of Lethbridge 

Department of Biology Lecture Series. Jan. 9. 

VP Palace, CL Baron, RE Evans KG Wautier, C Doebel, and S Kollar.2006. Mobility of arsenic in a freshwater 

system receiving historical gold mine tailings depositions. Risk assessment for native fishes with links to human 

health. Western College of Veterinary Medicine, Department of Verterinary Pathology lecture series. Feb. 6. 

Abstracts Presented: 

Carroll L, Thébeau N, Halden N, Hanson M, Palace V. 2011. A Retrospective Analysis of 

Manganese in Lake Trout (Salvelinus namaycush) Otoliths: links to reproductive failure. 32nd 
13

Annual Meeting of the Society of Environmental Toxicology and Chemistry, Boston MA, 

Nov. 13-17. 

Park B, Palace VP, Wautier K, Blanchfield P, Kidd K. 2011. Recovery of wild fish populations 

after a whole-lake estrogen addition. 32nd Annual Meeting of the Society of Environmental 

Toxicology and Chemistry, Boston MA, Nov 

Beaudry M, Mandrak N, CampbellR, GlebeB, Palace VP. 2011. Assessment of the potential 

effects of the aquatic herbicide, Diquat, on the growth of juvenile Spotted Gar (Lepisosteus 

oculatus). 21st Annual Meeting of the European Society of Environmental Toxicology and 

Chemistry, Milan, Italy. May 15-19. 

Fernie K, Letcher R, Alaee M, Champoux L, Brimble S, Palace VP. 2011. Changes in retinol 

and thyroid function of wild peregrine falcon (Falco peregrinus) nestlings relative to brominated 

flame retardants. 21st Annual Meeting of the European Society of Environmental Toxicology 

and Chemistry, Milan, Italy. May 15-19. 

Friedrich LA, Halden NM, Palace VP. 2011. Selenium uptake in otoliths from cold-water fish 

species captured downstream from coal mining. Goldschmidt Conference, Prague Czech 

republic August 14-19. 

Darling C, Palace VP, Tomy G. 2011 Thyroid hormone extraction from plasma: exploring pH 

and solvent choice. 31st International Symposium on Halogenated Persistent Organic Pollutants

POPs’ Science in the Heart of Europe. Brussels, Belgium. August 21-25. 

Fernie K, Champoux L, Brimble S, Alaee M, Palace VP. 2011. Spatial patterns and correlations 

of retinol, thyroid hormones and flame retardants in nestling Peregrine Flacons across the 

Canadian Great Lakes-St. Lawrence basin. 31st International Symposium on Halogenated 

Persistent Organic Pollutants POPs’ Science in the Heart of Europe. Brussels, Belgium. August 

21-25.

Park B, Palace V, Wautier K, Gemmill B, Tomy G. 2010. Impacts of dietary 

tetrabromoethylcyclohexane (TBECH) on the thyroid axis of brown trout (Salmo trutta). 31st 

Annual Society of Environmental Toxicology and Chemistry Meeting. Nov. 7-11, Portland, OR 

Palace VP, Park B, Wautier K, Gemmill B, Tomy G. 2010. Impact of dietary 

tetrabromoethylcyclohexane (TBECH) on the thyroid axis of brown trout (Salmo trutta). 30TH

International Dioxin Symposium, Sept. 12-17, San Antonio TX 

Sverko E, Tomy GT, Palace V, Smith LAP, McCarry BE. 2010. Fate of related compounds to 

dechlorane plus in a lake Ontario (Canada) foodweb. 30th International Dioxin Symposium, 

Sept. 12-17, San Antonio TX 

Tomy GT, Gemmill B, Pleskach K, Peters L, Palace V, Wautier K, Park B, Darling C, 

Rosenberg B, McCrindle R. 2010. Toxicokinetics of 1,2-dibromo-4-(1,2-
14

dibromoethyl)cyclohexane in juvenile brown trout (Salmo trutta) and effects on plasma sex 

hormones. 30th International Dioxin Symposium, Sept. 12-17, San Antonio TX.

Palace VP, Tomy G, Darling C, Fernie K, Stapleton HM. 2010. Development of an LC-MS/MS 

Based Assay for type I and II deiodianse enzyme activities. 20th Annual Meeting of the 

European Society of Environmental Toxicology and Chemistry, Seville, Spain. May 23-27. 

Friedrich L, Isinguzu I, Orr P, Arnett T, Halden, Palace VP. 2010. Selenium in annual growth 

zones of otoliths from three cold-water fish species captured downstream from coal mining 

activity. 20th Annual Meeting of the European Society of Environmental Toxicology and 

Chemistry, Seville, Spain. May 23-27. 

Hare J, Tomy G, Wautier K, Mittermuller S, StrugerJ, Palace VP. 2010. Survival, growth and 

reproduction in fathead minnows (Pimephales promelas) exposed to environmentally relevant 

concentrations of atrazine, chlorpyrifos and clopyralid. 20th Annual Meeting of the European 

Society of Environmental Toxicology and Chemistry, Seville, Spain. May 23-27. 

Gemmill B Tomy G, Pleskach K, Peters L, Palace V, Wautier K, Park B. 2010. 

Bioaccumulation, metabolism of 1,2-dibromo-4(1,2-dibromoethyl)-cyclohexane in juvenile 

brown trout and effects on circulating levels of testosterone. 20th Annual Society of 

Environmental Toxicology and Chemistry European Meeting, May 23-27, Seville, Spain. 

Palace VP, Blanchfield P, Tate L, Mills K, Park B, Wautier K, Evans RE, Kidd K. 2010. Effects 

of whole lake additions of a potent estrogen. 64th Canadian Conference for Fisheries Research 

Toronto, January 6-8. 

Godard DR, Groman DB, Park BJ, Palace VP. 2009. Pathological effects of sub-lethal explosive 

based instantaneous pressure change (IPC) on fish. 14th International Conference of the 

European Association of Fish Pathologists, Prague, CZ, September 14-19.

Godard DR, Park BJ, Palace VP. 2009. Histopathological assessment of the sub-lethal effects of 

explosive based instantaneous pressure changes (IPCs) on fish. Regional Habitat Coordinating 

Committee Meeting, Saskatoon, SK, January 13-14 

Kidd, K., Blanchfield, P., Mills, K., Tate, L., Palace, V. and Park, B. 2009. Is the birth control 

pill an effective form of contraception for wild fishes? 36th Annual Aquatic Toxicity Workshop, 

Charlevoix, Quebec. 27-30 September. 

Carroll LC, Thébeau N, Halden NM, Palace VP. 2009. Determining whether manganese 

exposure can be linked to reproductive failure in Lake Trout from Pipestone Bay, Ontario using 

LA-ICP-MS analysis of otoliths. Prairie University Biological Seminar. Lethbridge AB, Feb 19-

22 
15

Carroll LC, Thébeau N, Halden NM, Hanson M, Palace VP. 2009. Determining metal 

exposure in Lake Trout from Pipestone Bay, Ontario using LA-ICP-MS analysis of otoliths. 4th 

International Otolith Symposium, Monterey CA. Aug. 23-28. 

Carroll LC, Thébeau N, Halden NM, Palace VP. 2009. Incorporation of Cu, Cd, Ni, Pb and Zn in 

otoliths of rainbow trout (Oncorhynchus mykiss) fed experimental diets. 30th Annual Meeting of 

the Society of Environmental Toxicology and Chemistry. New Orleans LA, Feb. 19-23. 

Godard DR, Groman DB, Park BJ, Palace VP. 2009. Pathological assessment of explosive based 

instantaneous pressure chance (IPC) on fish. 62nd Canadian Conference for Fisheries Research 

Ottawa, January 10. 

Carroll LC, Thébeau N, Halden NM, Hanson M, Palace VP. 2009. Microchemical analysis of 

otoliths from lake trout. 62nd Canadian Conference for Fisheries Research 

Ottawa, January 10. 

Hare J, Tomy G, Wautier K, Mittermuller S, Palace VP. 2008. Assessment of potential growth 

and reproductive effects in fathead minnows (Pimephalus promelus) exposed to environmentally 

relevant waterborne concentrations of atrazine, glyphosate, clopyralid and chlorpyrifos. 29th

Annual Society of Environmental Toxicology and Chemistry Meeting, Tampa FL, Nov. 16-20. 

Carroll L, Thébeau N, Halden N, Palace VP. 2008. Examining links between manganese 

exposure and reproductive failure in lake trout using LA-ICP-MS analysis of archived otoliths. 

29th Annual Society of Environmental Toxicology and Chemistry Meeting, Tampa FL, Nov. 16-

20. 

Palace VP, ParkB, Pleskach K, GemmillB, Tomy G. 2008. Altered thyroid hormone metabolism 

in Rainbow trout exposed to HBCD Diastereoisomers. Dioxin 2008, Birmingham, UK. August 

17-22. 

Peters LE, Palace VP, Pleaskach K, Tomy GT. 2008. Examination of BDE-47 in vitro exposure 

on steroidogenesis in rainbow trout (Onchorhynchus mykiss) using LC/MS/MS. Dioxin 2008, 

Birmingham, UK. August 17-22. 

Baron, CL, Palace VP, Park B, Pleskach K, Gemmill B, Tomy G. 2008. Does oral exposure to 

Individual HBCD Diastereoisomers Alter thyroxine (T4) metabolism in Juvenile Rainbow trout? 

35Th aquatic Toxicity Workshop, Saskatoon, SK. Oct. 5-8. 

Miller, L, Rasmussen J, Palace V, Wang F, Carroll L, Hontela A. 2008. Physiological and 

biochemical responses of rainbow trout and brook trout exposed to elevated selenium from coal 

mines. 35Th aquatic Toxicity Workshop, Saskatoon, SK. Oct. 5-8. 

Hare J, Wautier K, Tomy G, Struger J, Mittermuller S, Palace VP. 2008. Potential growth and 

reproductive effects in fathead minnows (Pimephalus promelus) exposed to environmentally 

relevant waterborne concentrations of atrazine, glyphosate, clopyralid and chlorpyrifos. 29th 
16

Annual Meeting of the Society of Environmental Toxicology and Chemistry, Tampa FL, Nov. 

16-20.

Carroll L, Thébeau N, Halden N, Palace V. 2008. Examining links between manganese exposure 

and reproductive failure in lake trout using LA-ICP-MS analysis of archived otoliths. 29th 

Annual Meeting of the Society of Environmental Toxicology and Chemistry, Tampa FL, Nov. 

16-20. 

Orihel D, Muir D, Darling C, Dupuis A, Palace V, Park B, Paterson M, Tomy G. 2008. Fate of 

Decabromodiphenyl Ether Added to Mesocosms in a Boreal Lake: Preliminary Sediment 

Results. 29th Annual Meeting of the Society of Environmental Toxicology and Chemistry, 

Tampa FL, Nov. 16-20. 

Godard D, Hanson M, Palace VP. 2008. Pathological Assessment of Blood Biochemistry 

Following Sub-Lethal Explosive Based Instantaneous Pressure Changes (IPCs) on Fish. 30th Annual 

Meeting of the Society of Environmental Toxicology and Chemistry, New Orleans LA, Nov. 19-23. 

Carroll L, Thébeau N, Halden N, Palace V. 2009. Incorporation of Cu, Cd, Ni, Pb and Zn in 

otoliths of rainbow trout (Oncorhynchus mykiss) fed experimental diets. 30th Annual Meeting of 

the Society of Environmental Toxicology and Chemistry, New Orleans LA, Nov. 19-23.

Kidd, K., P. Blanchfield, K. Mills, M. Paterson, C. Podemski, D. Findlay, V. Palace. 2007. 

Exposure of aquatic populations to a potent estrogen: what are the risks? 28th Annual Meeting of 

the Society of Environmental Toxicology and Chemistry, November 11-15, Milwaukee WI. 

Palace, VP, C. LeVasseur, N. Halden, and G. Sterling 2007. Selenium exposure histories 

derived from analysis of selenium in otoliths of rainbow trout (Oncorhynchus mykiss) and brook 

trout (Salvelinus fontinalis) from an area impacted by coal mining In Alberta’s eastern slopes. 

28th Annual Meeting of the Society of Environmental Toxicology and Chemistry, November 11-

15, Milwaukee WI. 

Peters, L., K. Pleskach, V.P. Palace, G.T. Tomy. 2007. Does BFR 47 affect in vitro 

steroidogenesis in fish gonadal tissue? 28th Annual Meeting of the Society of Environmental 

Toxicology and Chemistry, November 11-15, Milwaukee WI. 

Palace, VP., S. Mittermuller, C.L. Baron, A.T. Goodmanson, J. Hare, K.G. Wautier, J. Struger, 

L.E. Peters. 2007. Site fidelity of golden shiners (Notemigonus crysoleucas) and 

their use in effects-based assessments of pesticides by DFO’s Center for 

Environmental Research on Pesticides (CERP). 28th Annual Meeting of the Society of 

Environmental Toxicology and Chemistry, November 11-15, Milwaukee WI. 

Palace, VP., B. Park, L.E. Peters and C.L. Baron. 2007. Effects of environmentally relevant 

concentrations of atrazine on growth and gonad development in fathead minnows (Pimephales 
17

promelas). 28th Annual Meeting of the Society of Environmental Toxicology and Chemistry, 

November 11-15, Milwaukee WI. 

Palace, VP., K. Pleskach, K. Wautier, G. Tomy. 2007. Effects of Hexabromocyclododecane 

(HBCD) diastereoisomers on the thyroid axis of Juvenile Rainbow trout (Oncorhynchus mykiss). 

28th Annual Meeting of the Society of Environmental Toxicology and Chemistry, November 11-

15, Milwaukee WI. 

Bartlett, A, J. Struger and V. P. Palace. 2007. Use of acetylcholinesterase as an indicator of 

pesticide exposure from in situ studies using the freshwater amphipod Hyalella azteca. 34th

Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, Halifax, NS. 

Miller, L., J. Rasmussen, V. Palace, F. Wang and A. Hontela. 2007. The effects of agricultural 

drain water on the physiological stress response in fish - interactions and effects of multiple 

chemical stressors. 34th Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, Halifax, NS. 

Fisk, A., A. Buckman, V. Palace and S. Brown. 2007. Factors influencing the bioaccumulation 

of PCBs by fish and potential interactions with the thyroid axis. 34th Annual Aquatic Toxicity 

Workshop, Sept 30-Oct. 3, Halifax, NS. 

Palace, V., R. Evans, C. Baron and A. Fisk. 2007. Scott Brown's contribution to our 

understanding of how PCBs affect vitamin metabolism. 34th Annual Aquatic Toxicity Workshop, 

Sept 30-Oct. 3, Halifax, NS. 

LeVasseur, C, G. Sterling, N. Halden and V. Palace. 2007. Selenium exposure histories derived 

from analysis of selenium in otoliths of rainbow trout (Oncorhynchus mykiss) and brook trout 

(Salvelinus fontinalis) from an area impacted by coal mining In Alberta’s eastern slopes. 34th

Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, Halifax, NS. 

Palace, V.,, A. Goodmanson, J. Struger, S. Mittermuller, J. Hare, K. Wautier, C. 

Baron and L. Peters. 2007. Site fidelity of golden shiners (Notemegonus chrysoleucas) and their 

use in effects-based assessments of pesticides by DFO’s Center for Environmental 

Research on Pesticides (CERP). 34th Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, 

Halifax, NS. 

Palace, VP., K. Pleskach, K. Wautier, G. Tomy. 2007. Effects of Hexabromocyclododecane 

(HBCD) diastereoisomers on the thyroid axis of Juvenile Rainbow trout (Oncorhynchus mykiss). 

34th Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, Halifax, NS. 

Palace, VP., L. Peters, P. Cott, K. Wautier, R. Evans, B. Hanna and D. Godard. 2007. 

Development of quantitative histopathology tools to assess instantaneous 

pressure change induced effects in early life stages of rainbow trout (Onchorhynchus mykiss). 

34th Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, Halifax, NS. 
18

Peters, L., K. Pleskach, V.P. Palace, G.T. Tomy. 2007. Does BFR 47 affect in vitro 

steroidogenesis in fish gonadal tissue? 34th Annual Aquatic Toxicity Workshop, Sept 30-Oct. 3, 

Halifax, NS. 

Palace VP, K Pleskach, T Halldorson, R Danell, K Wautier, R Evans, M Alaee, C Marvin and G 

Tomy. 2007. Biochemical effects of hexabromocyclododecane (HBCD) diastereoisomers in 

juvenile rainbow trout (Onchorhynchus mykiss). 4th International workshop on Brominated 

Flame Retardants. Amsterdam, Netherlands. April 24-27. 

Tomy G, VP Palace K Pleskach, R Dannell and T Halldorson. 2007. Dietary exposure of 

juvenile rainbow trout (Onchorhynchus mykiss) to 1,2-bis(2,4,6-tribromophenoxyethane) 

(BTBPE): bioaccumulation parameters and biochemical effects. 4th International workshop on 

Brominated Flame Retardants. Amsterdam, Netherlands. April 24-27. 

Tallman, RT, RA Bodaly, KL Howland, KH Mills, VP Palace. 2007. Impacts of industrial 

development on coregonid populations of northern Canada. Proceedings of the XXth 

International Conference on Coregonids, Slovenia. 

Peters L, V Palace, P Cott, D Godard, K Wautier, RE Evans, D Clayton, A Wong, D Monita, B Hanna, W 

Morris, D Wright, K Bill. 2006. Examination of the Potential Effects of Instantaneous Pressure Changes on 

Eggs, Sac Fry and Juvenile Rainbow Trout (Onchorhynchus mykiss). 30th Annual Larval Fish Conference, 

Lake Placid NY, Sept. 10-14.

Fudge T, K Wautier, G Sterling and VP Palace. 2006. Rainbow trout fry emergence from 

artificial redds containing different fine sediment loads. 30th Annual Larval Fish Conference, Lake 

Placid NY, Sept. 10-14.

Bartlett AJ, Struger J, and VP Palace. 2006. In situ exposure of a freshwater invertebrate, 

Hyalella azteca, to pesticide-impacted tributaries of Lake Ontario. 27th Annual Society of 

Toxicology and Chemistry Meeting, Montreal, QC. Nov. 5-9. 

Palace VP, J Struger, B Park, K Wautier, R Evans, L Peters, C Baron. 2006. Effects-based 

assessments of pesticides on wild fish populations by DFO’s center for Environmental Research 

on Pesticides (CERP). 27th Annual Society of Toxicology and Chemistry Meeting, Montreal, 

QC. Nov. 5-9. 

 

L Peters, V Palace and G Tomy. 2006. Development of a non-lethal method to determine sex 

steroid hormone levels in fish using fecal extraction. 27th Annual Society of Toxicology and 

Chemistry Meeting, Montreal, QC. Nov. 5-9. 

VP Palace, GS Sterling, P Siwik, RE Evans, N Halden, KG Wautier, J Holm. 2006. Limitations 

of non-lethal sampling for determining spatiotemporal exposure to selenium in fish from mine 

impacted sites. 27th Annual Society of Toxicology and Chemistry Meeting, Montreal, QC. Nov. 

5-9. 

VP. Palace, LE. Peters, Erida, Marcos, S Kollar, A Bartlett, M Gray, J Werner J Emmert and Y 
19

Carolsfeld.2006. Investigating potential causes of episodic mortality in culturally and 

commercially important fish from the Sao Francisco River, Brazil, using a small bodied fish 

species. 27th Annual Society of Toxicology and Chemistry Meeting, Montreal, QC. Nov. 5-9. 

 

VP Palace, N Halden, R Evans, G Sterling, J Holm, P Siwik. 2006. Determining life histories of 

fish using laser ablation inductively coupled plasma mass spectrometric analysis (LA-ICP-MS) 

of selenium in otoliths. 27th Annual Society of Toxicology and Chemistry Meeting, Montreal, 

QC. Nov. 5-9. 

Palace, VP, Peters L, Cott P, Godard D, Wautier K, Evans R, Clayton D, Wong A, Monita D, 

Hanna B. 2006. Examination of the Potential Effects of Instantaneous Pressure Changes on 

Eggs, Sac Fry and Juvenile Rainbow Trout (Onchorhynchus mykiss) 27th Annual Society of 

Toxicology and Chemistry Meeting, Montreal, QC. Nov. 5-9. 

Palace V, Kidd K, Blanchfield P, Mills K, Wautier K, Lazorchak J, Flick R. 2006. A 

comparison of subcellular- through population-level effects of a potent estrogen mimic on fish: 

results from a multi-year whole-lake experiment. 27th Annual Society of Toxicology and 

Chemistry Meeting, Montreal, QC. Nov. 5-9. 

VP Palace, J Struger, B Park, K Wautier, R Evans, L Peters, C Baron. 2006. Effects-based 

assessments of pesticides on wild fish populations by the Canadian Department Fisheries and 

Oceans’ Center for Environmental Research on Pesticides (CERP). ANCAP / SETAC Africa 

International Conference on Pesticide Use in Developing Countries 

Environmental Fate, effects and Public Health Implications.

16 - 20 October 2006, Arusha, Tanzania. 

Peters, L, VP Palace, G Tomy. 2006. Development of non-lethal methods to determine 

reproductive potential in fish using fecal extractions. ANCAP / SETAC Africa International 

Conference on Pesticide Use in Developing Countries Environmental Fate, effects and Public 

Health Implications. 16 - 20 October 2006, Arusha, Tanzania. 

Parsons, MB, PK Smith, TA Goodwin, GEM Hall, JB Pecival, R Mroz, KG Doe, K Tay, VP 

Palace. 2006. Assessing and reducing risks from historical gold mine tailings in Nova Scotia, 

Canada. 7Th International Symposium on Environmental Geochemistry. Beijing, China. Sept. 24-

27. 

Peters LE, Struger J, Palace VP. 2006. Effects Based Assessment of Golden Shiners 

(Notemigonus crysoleucas) Collected From a Lake Ontario Tributary Surrounded by Pesticide 

Usage. 59th Canadian Conference for Fisheries Research, Calgary, Alberta, 5-7 January. 

Miller LL, J Rasmussen, V Palace, A Hontela. 2006. The effect of Se on the physiological stress 

response of rainbow trout; a field study. 59th Canadian Conference for Fisheries Research, 

Calgary, Alberta, 5-7 January. 
20

Miller, LL, JB Rasmussen, V Palace, A Hontela. 2005. The effects of selenium on the 

physiological stress response of rainbow and brook trout from a coal mining area. Canadian 

Society for Zoology. Kingston, Ontario. May 5-7. 

Palace, VP and L.E. Peters. 2005. Sublethal Toxicity Examinations and Environmental Effects 

Monitoring. Reducao de Impactos-Ambientais na Bacia do Alto Medio Rio Sao Francisco: 

Forum de Coopercao Interninstitucional em Pesquisas e Acores. September 29-30, Tres Marias 

Brazil. 

Struger, J., Sverko, E., Marvin, C., Dove, A., Backus, S., Muir, D., Martin, P., Hewitt, M., 

Thompson, D., Palace, V. 2005. Occurrence of current-use pesticides in surface waters of 

Ontario, Canada. 26th Annual Meeting of the Society of environmental Toxicology and 

Chemistry, Baltimore MA, Nov. 13-17. 

Peters LE, Munkittrick KR, Park BJ, Struger J, Palace VP. 2005. In Vitro Steroidogenesis in 

Golden Shiner (Notemigonus crysoleucas) Collected From a Lake Ontario Tributary Surrounded 

by Pesticide Usage. 26th Annual Meeting of the Society of environmental Toxicology and 

Chemistry, Baltimore MA, Nov. 13-17. 

Palace VP, Cott P, Godard D, Wautier K, Clayton D, Wong A, Monita D, Hanna B, Morris W, Wright D, 

Bill K. 2005. Examination of the potential effects of Instantaneous pressure changes on larval rainbow 

trout (Onchorhynchus mykiss). 26th Annual Meeting of the Society of environmental Toxicology 

and Chemistry, Baltimore MA, Nov. 13-17. 

Podemski C, S Kollar, VP Palace. 2005. Selenium in northern stream foodwebs. Selenium 

Science Workshop. June 28-29 Hinton, AB. 

Palace VP, J Holm, K Wautier, RE Evans, CL Baron, P Siwik, G. Sterling. 2005. An 

Examination of Se Induced Deformities in Salmonids Captured Near Coal Mining Activity 

in Alberta's Northeast Slopes Region. Selenium Science Workshop. June 28-29 Hinton, AB. 

Doebel C, Baron CL, Evans RE, Palace VP. 2005. The use of small bodied forage fish in caged 

exposures for metal mining environmental effects monitoring. Proceedings of the International 

Conference on Heavy Meals in the Environment. Rio DeJaneiro, Brazil. June 5-9, 2005. 

Gray M, Munkittrick K, Dahl J, Baron C, Wautier K, Kollar S, Werner J, Palace V. 2005. 

METALS and metallothionein in slimy sculpin (Cottus cognatus) captured at a Diamond mining 

Site in Northern Canada. Proceedings of the International Conference on Heavy Meals in the 

Environment. Rio DeJaneiro, Brazil. June 5-9, 2005. 

Palace VP, Park B, Struger J, Parrott J, Evans R, Wautier K, Baron C, Kidd K, Gray, M. 2004. 

An initial investigation of the potential impacts of pesticides on Golden Shiners in a Niagara 

region stream. 25th Annual Society of Environmental Toxicology and Chemistry Meeting, 

Portland, OR, Nov. 14-18. 
21

 

Law, K, Halldorson, T, Danell, R, Palace, V, Wautier, K, Evans, B, Whittle, M, Alaee, M, 

Marvin, C, Tomy, G. 2004. Evidence of bioisomerization of - and -hexabromocyclododecane 

(HBCD) isomers in fish. 25th Annual Society of Environmental Toxicology and Chemistry 

Meeting, Portland, OR, Nov. 14-18. 

Law, K, Palace, V, Halldorson, T, Danell, R, Wautier, K, Evans, B, Brinkworth, L, Alaee, M, 

Tomy, G. 2004. Dietary accumulation of hexabromocyclododecane (HBCD) isomers in juvenile 

rainbow trout (Oncorhynchus mykiss). 25th Annual Society of Environmental Toxicology and 

Chemistry Meeting, Portland, OR, Nov. 14-18. 

 

Doebel, C, Baron, C, Wautier, K, Evans, R, Palace, V. 2004. The Use of Small Bodied Forage 

Fish in Caged Exposures for Metal Mining Environmental Effects Monitoring (EEM). 25th 

Annual Society of Environmental Toxicology and Chemistry Meeting, Portland, OR, Nov. 14-

18. 

Lazorchak, J, Palace, V, Evans, R, Wautier, K, Flick, R, Wiechman, B, Braam, A, Kidd, K, 

Lattier, D. 2004. Molecular, Protein and Histological Indicators in Fathead Minnow and Pearl 

Dace Exposed to EE2 in a Whole Lake Dosing Experiment . 25th Annual Society of 

Environmental Toxicology and Chemistry Meeting, Portland, OR, Nov. 14-18. 

Kidd, K, Palace, V, Blanchfield, P, Mills, K, Wautier, K, Lazorchak, J, Lattier, D. 2004. 

Differences in the biochemical- through population-level responses of two minnow species 

exposed to a potent estrogen mimic. . 25th Annual Society of Environmental Toxicology and 

Chemistry Meeting, Portland, OR, Nov. 14-18. 

Martin, P, Mayne, G, Bursian, S, Palace, V, Tomy, G. 2004. Altered Thyroid Status and Vitamin 

A Levels in Mink (Mustela vison) Exposed to a Commercial PBDE Mixture. 25th Annual 

Society of Environmental Toxicology and Chemistry Meeting, Portland, OR, Nov. 14-18. 

Martin, P, Bursian, S, Mayne, G, Tomy, G, Palace, V. 2004. Immunomodulation in Ranch Mink 

(Mustela vison) Induced by Dietary Exposure to a Commercial PBDE Mixture. 25th Annual 

Society of Environmental Toxicology and Chemistry Meeting, Portland, OR, Nov. 14-18. 

Doebel C, CL Baron, KG Wautier, RE Evans and VP Palace. 2004. The Use of Small Bodied 

Forage Fish in Caged Exposures For Metal Mining Environmental Effects Monitoring (EEM). 

31st Annual Aquatic Toxicity workshop, Charlottetown, PEI. Oct. 24-27. 

Palace VP, Doebel C, CL Baron, KG Wautier, RE Evans. 2004. Caged, Small Bodied Forage 

Fish For Metal Mining Environmental Effects Monitoring (EEM). 31st Annual Aquatic Toxicity 

workshop, Charlottetown, PEI. Oct. 24-27. 

Gray M, C Smith, B Park, V Palace, K Kidd. 2004. Depression of brain acetylcholinesterase 

(AChE) in fish collected downstream of agricultural activities. 31st Annual Aquatic Toxicity 
22

workshop, Charlottetown, PEI. Oct. 24-27. 

Palace VP, RE Evans, KG Wautier, B Park, CL Baron, K Kidd, J Struger, J Parrott. 2004. 

Initial Examinations of the Potential for Pesticides to Impact Golden Shiners in a Niagara 

Stream. 31st Annual Aquatic Toxicity workshop, Charlottetown, PEI. Oct. 24-27. (Invited paper) 

Parsons M, P Smith, T Goodwin, GE Hall, S Winch, V Palace, R Parrott, A Sangster. 2004. 

Contamination of freshwater and marine environments from historical gold mining activities in 

Nova Scotia. 31st Annual Aquatic Toxicity workshop, Charlottetown, PEI. Oct. 24-27. 

VP Palace, RE Evans, KG Wautier, CL Baron, J Holm. 2004.Toxicity of Selenium to 

Freshwater Fishes. Atelier Workshop on Selenium: Selenium and hydroelectric reservoirs. 

March 22, Montreal, QC. 

V.P. Palace, C. Doebel, C.L. Baron, R.E. Evans, K.G. Wautier, 

J. Werner, S. Kollar and J.F. Klaverkamp. 2004. The potential for using caged small bodied fish 

for metal mining environmental effects monitoring (EEM). Environmental Effects Monitoring 

Science Symposium, Feb. 16-18, Fredericton NB. 

Martin PA, Bursian SJ, Palace VP, Mayne G. 2004. Thyroid and vitamin A status in mink fed 

PCB-contaminated carp from Saginaw Bay. 24th Annual Society of Toxicology and chemistry 

Meeting, November 9-13, Austin TX. 

Palace VP. 2004. The Department of Fisheries & Oceans’ Center for Environmental Research 

On pesticides: program overview. Federal Prairie Water Workshop, Feb. 24-25, Regina 

Saskatchewan. 

Flick R, Lattier D, Kidd K, Palace V, Smith M, Wiechman B, Hodge T, Toth G. 2004. 

Vitellogenin gene expression in fathead minnows exposed to EE2 in a whole lake dosing 

experiment. 24th Annual Society of Toxicology and chemistry Meeting, November 9-13, Austin 

TX. 

Palace VP, Kidd K, Blanchfield P, Evans B, Wautier K. 2004. Vitellogenin induction and 

histopahtological effects in pearl dace from a lake experimentally treated with synthetic 

estrogen.. 24th Annual Society of Toxicology and chemistry Meeting, November 9-13, Austin 

TX. 

Werner J, Palace VP, Baron CL, Evans RE, Wautier KG, Werner JP, Kollar S, Klaverkamp JF. 

2004. Metals, metalloids and metallothionein in tissues of fish from a Canadian freshwater 

system receiving gold mining effluents. 24th Annual Society of Toxicology and chemistry 

Meeting, November 9-13, Austin TX. 

Palace VP, Werener J, Wautier KG, Baron CL, Evans RE, Cobb DG. 2004. Lipid soluble 

vitamins A and E as integrative measures of fish health for environmental effects monitoring 

(EEM). 24th Annual Society of Toxicology and chemistry Meeting, November 9-13, Austin TX. 
23

Palace VP, Baron CL, Evans RE, Wautier KG, Brinkworth L, Kollar S, Werner JP. 2004. The 

mobility of arsenic in a Canadian freshwater system receiving gold mining effluents. 24th Annual 

Society of Toxicology and chemistry Meeting, November 9-13, Austin TX. 

Blanchfield P, Majewski A, Palace VP, Kidd K, Mills K. 2004. Reproductive and popul;ation 

level impacts of a synthetic estrogen on fathead minnows. 24th Annual Society of Toxicology 

and chemistry Meeting, November 9-13, Austin TX. 

V.P. Palace, C. Doebel, C.L. Baron, R.E. Evans, K.G. Wautier, J. Werner, S. Kollar and J.F. 

Klaverkamp. 2004. The Potential for Using Caged Small Bodied Fish For Metal Mining 

Environmental Effects Monitoring (EEM). EEM Science Symposium, Fredericton, New 

Brunswick, Feb. 16-18. 

V.P. Palace, J.Werner, K. Wautier, C.L. Baron, R.E. Evans and D. Cobb.2003. Lipid Soluble 

Vitamins A and E as Integrative Measures of Fish Health for EEM. Offshore Oil and Gas 

Environmental Effects Monitoring Workshop:Approaches and Technologies. Bedford Institute 

of Oceanography, Nova Scotia, Canada, May 26-30, 2003 

Palace, VP, CL Baron, RE Evans, K Wautier. 2003. Mobility of Arsenic in A Canadian 

Freshwater System Receiving Gold Mine Effluents. Proceedings of the XIIth conference on 

Heavy Metals in the Environment. Grenoble, France. May 26-30. 

Palace, VP, CL Baron, RE Evans, K Wautier, JF Klaverkamp. 2003. Metals, Metalloids and 

Metallothionein in Tissues of Fish from A Canadian Freshwater System Receiving Gold Mining 

Effluents and in Fish from Nearby Reference Areas. Proceedings of the XIIth conference on 

Heavy Metals in the Environment. Grenoble, France. May 26-30. 

Palace, VP, D Nicholas, F Rosenzweig, CL Baron, RE Evans, K Wautier, L Brinkworth. 2003. 

An Investigation into Microbially-Mediated Mobilization and Sequestration of Arsenic in 

Arsenic-Contaminated Sediments of Balmer Lake, Ontario. Metals in the Environment Research 

Network (MITE-RN) Symposium. Ottawa, ON. Feb 24-25. 

Palace, VP. 2003. Vitamin E as a mediator and indicator of toxicant induced stress in fish. 

Aquaculture America 2003, Louisville KY. Feb 18-21. (Invited Paper).

Klaverkamp, JF, VP Palace, CL Baron, RE Evans, KG Wautier. 2002. Cumulative impacts on 

caged fish of gold mine and municipal wastewater effluents and of garbage-disposal run-off. 23rd

Annual Society of Environmental Toxicology and Chemistry Meeting. Salt Lake City, UT. Nov. 

16-20.

Palace,VP, K Kidd, P Blanchfield, RE Evans, CL Baron, K Wautier. 2002. Reproductive effects 

in fish from a lake experimentally treated with the synthetic estrogen ethynylestradiol. 23rd

Annual Society of Environmental Toxicology and Chemistry Meeting. Salt Lake City, UT. Nov. 

16-20. 
24

Palace, VP, RJ Letcher, J Li, T Halldorson, E Braekevelt, R Danell, K Wautier, A Fisk, G 

Tomy. 2002. Dietary accumulation and biochemical effects of PBDEs in juvenile lake trout. 23rd

Annual Society of Environmental Toxicology and Chemistry Meeting. Salt Lake City, UT. Nov. 

16-20. 

Letcher, RJ, J Li, T Halldorson, E Braekevelt, R Danell, VP Palace, K Wautier, A Fisk, G 

Tomy. 2002. Dietary PBDE Exposure and Phenolic Metabolites in Juvenile Lake Trout (Salvelinus

namaycush). 23rd Annual Society of Environmental Toxicology and Chemistry Meeting. Salt 

Lake City, UT. Nov. 16-20. 

 

Lazorchak, JM, R Flick, DL Lattier, K Kidd, VP Palace, T Hodge. 2002. Vitellogenin Gene 

Expression in Fathead Minnows Exposed to EE2 in a Whole Lake Dosing Experiment. 23rd

Annual Society of Environmental Toxicology and Chemistry Meeting. Salt Lake City, UT. Nov. 

16-20. 

Tomy, G, VP Palace, E Braekevelt, S Tittlemier, B Lau, G Eales, J Plohman. 2002. 

Biotransformation of perfluoroooctane sulfonamide by liver microsomes in vitro. 23rd Annual 

Society of Environmental Toxicology and Chemistry Meeting. Salt Lake City, UT. Nov. 16-20. 

Palace, VP, J Holm, K Wautier, RE Evans, CL Baron, G Sterling, P Siwik, JE Spallholz. 2002. 

Selenium exposure and effects in wild rainbow trout and brook trout from Alberta's Northeast 

Slopes region. 23rd Annual Society of Environmental Toxicology and Chemistry Meeting. Salt 

Lake City, UT. Nov. 16-20. (Invited Paper)

Palace VP, K Kidd, P Blanchfield, RE Evans, K Wautier. 2002. Reproductive effects in wild 

fathead minnows from a lake experimentally treated with the synthetic estrogen, 

ethynylestradiol. 26th Annual Larval Fish Conference, Bergen, Norway. July 22-26. 

Palace, VP, J Holm, RE Evans, K Wautier, CL Baron, P Siwik, G Sterling. 2002. Incidence of 

Deformities in brook trout (Salvelinus fontinalis) and rainbow trout (Onchorhynchus mykiss) 

exposed to elevated selenium downstream from coal mining activity in the northeastern slopes 

region of Alberta, Canada. 26th Annual Larval Fish Conference, Bergen, Norway. July 22-26. 

K Kidd, C Podemski, M Paterson, D Findlay, A Salki, VP Palace, B Evans, P Blanchfield, K 

Mills, B Park, J Werner, S Brown, M McMaster, G Van Der Kraak, J Lazorchak, G Toth, D 

Lattier, M Bagley, K Liber, D Graham, K Londry. 2001. A Whole-Lake Experiment to Examine 

the Effects of a Synthetic Estrogen on Aquatic Populations. Toxic Substances Research Initiative 

Regional Conference 9-10 November, Hamilton, ON. 

Holm, J, RE Evans, K Wautier, CL Baron, P Siwik, G Sterling, VP Palace. 2001. Incidence of 

deformities in brook trout (Salvelinus fontinalis) exposed to elevated Se downstream from coal 

mining activity in Alberta’s northeast slopes region. 28th Annual Toxicity Workshop, Winnipeg, 
25

MB Sept. 30-Oct. 2. 

Kidd, K, VP Palace, P Blanchfield, K Mills, B Evans, M McMaster, SB Brown, G van Der 

Kraak, D Lattier, J Lazorchak. 2001. A whole-lake experiment to examine the effects of a 

synthetic estrogen on fish populations. Aquatic Toxicity Workshop, 30 September - 3 Oct., 

Winnipeg, MB 

Kidd, K, C Podemski, M Paterson, D Findlay, A Salki, VP Palace, B Evans, P Blanchfield, K 

Mills, B Park, J Werner, S Brown, M McMaster, G Van Der Kraak, J Lazorchak, G Toth, D 

Lattier, M Bagley, K Liber, D Graham, K Londry. 2001. A Whole-Lake Experiment to Examine 

the Effects of a Synthetic Estrogen on Aquatic Populations. Endocrine TIP Review Meeting, 

American Chemistry Council, 30 Oct., Chicago, IL. 

Kidd, K, VP Palace, RE Evans, K Mills, P Blanchfield, L Vandenbyllardt, K Wautier, J 

Lazorchak, D Lattier, R Flick. 2002. Biochemical and physiological effects in wild fathead 

minnows from a lake experimentally treated with the synthetic estrogen, ethynylestradiol. 

Canadian Symposium on Water Quality, 4-5 February, Burlington, ON. 

Kidd, K, C Podemski, M Paterson, D Findlay, A Salki, VP Palace, B Evans, P Blanchfield, K 

Mills, B Park, J Werner, SB Brown, M McMaster, G van Der Kraak, J Lazorchak, G Toth, D 

Lattier, M Bagley, K Liber, D Graham, K Londry. 2002. A Whole-Lake Experiment to Examine 

the Effects of a Synthetic Estrogen on Aquatic Populations. Canadian Society of Limnologists 

Meeting, 3-5 January, Vancouver, BC

Baron, CL, JF Klaverkamp, RE Evans, K Wautier, VP Palace. 2002. Applying environmental 

effects monitoring strategies and mitigation approaches to gold mines in Northwestern Ontario. 

1st national DFO science Workshop, Sidney, B.C. Nov. 14-16. 

Kidd, K, VP Palace, P Blanchfield, K Mills, RE Evans, M McMaster, S Brown, G van Der 

Kraak, D Lattier, J Lazorchak. 2001. A whole lake experiment to examine the effects of a 

synthetic estrogen on fish populations. 28th Aquatic Toxicity Workshop, Winnipeg, MB. Oct. 30-

Sept. 3. 

Danelisen, I, VP Palace, H Lou, PK Singal. 2001. Maintenance and utilization of retinol 

(Vitamin A) in adriamycin-induced cardiomyopathy. J of Mol and Cell Cardiol, 33: A25. 

Kumar, D, VP Palace, I Danelisen, BI Jugdutt, PK Singal. 2001. Probucol induced antioxidants 

confers protection against I-R injury. J of Mol and Cell Cardiol, 33: A63, 2001. 

Majewski, A, P Blanchfield, K Wautier, VP Palace. 2001. Environmentally relevant doses of 

ethynylestradiol affect reproductive behaviour in male fathead minnows (Pimephales promelas). 

28th Aquatic Toxicity Workshop, Winnipeg, MB. Oct. 30-Sept. 3. 

Palace, VP, K Kidd, K Wautier, RE Evans, TA Dick, J Werner, CL Baron. 2001. Freshwater 

fish exposed to environmental estrogen contaminants have altered lipid soluble vitamin status. 
26

28th Aquatic Toxicity Workshop, Winnipeg, MB. Oct. 30-Sept. 3. 

Holm, J, RE Evans, K Wautier, CL Baron, P Siwik, G Sterling, VP Palace. 2001. Evaluations of 

the reproductive fitness of wild rainbow trout and brook trout exposed to elevated selenium 

concentrations in an area of active coal mining. 22nd Annual Society of Environmental 

Toxicology and Chemistry Conference, Baltimore, MD, Nov. 11-15. 

Palace, VP, K Kidd, K Wautier, RE Evans, TA Dick, J Werner, C.L. Baron. 2001. Freshwater 

fish exposed to environmental estrogens have altered lipid soluble vitamin status. 22nd Annual 

Society of Environmental Toxicology and Chemistry Conference, Baltimore, MD, Nov. 11-15. 

Holm, J, VP Palace, P Siwik, G Sterling, RE Evans, CL Baron. 2001. Potential effects of 

selenium on the reproductive fitness of rainbow trout downstream from a coal mine endpit lake. 

Proceedings of the 40th Annual Meeting of the Society of Toxicology, San Francisco, CA. 

March 25-29. 

Palace, VP, RE Evans, K Kidd, TA Dick. 2001. Long term effects to somatic growth and gondal 

maturation in yellow perch (Perca flavescens) exposed for short durations to waterborne 

ethynylestradiol. Proceedings of the 40th Annual Meeting of the Society of Toxicology, San 

Francisco, CA. March 25-29. 

Palace, VP, JP Werner, RE Evans, K Wautier, CL Baron, K Kidd. 2000. Effects of waterborne 

ethynylestradiol exposure on metallothionein and lipid soluble vitamins in lake trout. 21st 

Annual Society of Environmental Toxicology and Chemistry Conference, Nashville, TN, Nov. 

12-16. 

Danelisen, I, T Li, VP Palace, PK Singal. 2000. Vitamin A metabolism in adriamycin-induced 

cardiomyopathy. XXII Annual Meeting of the American Society of the International Society for 

Heart Research. Louisville, KY. June 14-18. J. Mol. Cell. Cardiol. 32(5): A47. 

Palace, VP, TA Dick, K Wautier, CL Baron, RE Evans, JF Klaverkamp. 1999. Effects of 17-

estradiol exposure on metallothionein and fat soluble antioxidant vitamins in juvenile lake trout. 

26th Annual Aquatic Toxicity Workshop, Edmonton, AB, Oct. 2-6. 

Palace, VP, K Wautier, CL Baron, RE Evans, JF Klaverkamp. 1999. Effects of estradiol 

exposure on metallothionein and fat solubel vitamins in juvenile lake trout. 26th annual Aquatic 

Toxicity Workshop, Edmonton, AB, Oct. 2-6.

Palace, VP, MF Hill, N Khaper, PK Singal. 1999. The heart consumes more vitamin A after a 

myocardial infarction. Experimental Biology, Washington D.C., April 17-21.

Palace, VP, M Hill, F Farahmand and PK Singal. 1998. Antioxidant vitamin pools and 

hemodynamic function following myocardial infarction in rats. XXth Annual Meeting of the 

International Society for Heart Research-American Section, Ann Arbor MI, August 9-12. J. Mol. 

Cell. Cardiol. 30(7):A239 
27

Kumar D, L Kirshenbaum, T Li, V Palace, and PK Singal. 1998.Apoptotic and necrotic cell 

death in adriamycin cardiomyopathy. XXth Annual Meeting of the International Society for 

Heart Research-American Section, Ann Arbor MI, August 9-12. J. Mol. Cell. Cardiol. 

30(7):A103. 

Palace, VP, D Kumar, MF Hill, PK Singal. 1998. Regional distribution of non-enzymatic 

antioxidants and oxidative stress in isolated perfused rat hearts. XVI World Congress of the 

International Society for Heart Research, Rhodes, Greece, May 27-31. 

Palace, VP, D Kumar, N Khaper, PK Singal. 1998. Oxidative stress induced regional changes in 

non-enzymatic antioxidants, redox ratio and lipid hydroperoxides in the heart. Experimental 

Biology, San Francisco, April 18-22. FASEB J. 12(4): A70. 

Kumar, D, L Kirshenbaum, T Li, VP Palace, PK Singal. 1998. Apoptotic and necrotic cell death 

in adriamycin cardiomyopathy. XVI World Congress of the International Society for Heart 

Research, Rhodes, Greece, May 27-31. 

Pichardo, J, VP Palace, N Khaper, PK Singal. 1997. Myocardial oxidative stress changes during 

compensated right heart failure in rats. Fifth World Congress of the International Society for 

Adaptive Medicine. Framingham, MA. September 7-10. 

Palace, VP, D Kumar, J Pichardo, N Khaper, PK Singal. 1997. Ischemia-reperfusion induced 

changes in non-enzymatic antioxidants and oxidative stress in different regions of the isolated rat 

heart. Second World Congress of the International Society of Molecular Nutrition and Therapy. 

Winnipeg, MB. August 2-4. 

Palace, VP, D Kumar, MF Hill, PK Singal. 1997. Regional distributions of non-enzymatics 

antioxidants and oxidative stress in isolated perfused rat hearts. Second World Congress of the 

International Society of Molecular Nutrition and Therapy. Winnipeg, MB. August 2-4. 

Palace VP, SB Brown, CL Baron, J Fitzsimons, JF Klaverkamp. 1996. Oxidative stress indices 

in adult lake trout (Salvelinus namaycush) from Lake Ontario and Lake Superior. American 

Fisheries Society Early Mortality Special Symposium. Dearborn, MI. August 26-29. Invited 

Paper. 

Brown, SB, VP Palace, JF Klaverkamp, J Fitzsimons. 1996. Vitamin dynamics in developing 

lake trout (Salvelinus namaycush) from Lake Ontario. American Fisheries Society Early 

Mortality Special Symposium. Dearborn, MI. August 26-29. Invited Paper. 

Palace VP, SB Brown, CL Baron, JF Klaverkamp. 1996. Insights into retinoid metabolism 

provided by recoveries of 3

H-retinol from tissues of lake trout (Salvelinus namaycush) preexposed to PCB 126. International Association for Great Lakes Research.Erindale College, 
28

Mississauga, ON. May 25-30. Invited Paper. 

Brown, SB, VP Palace. 1996 Vitamin dynamics in fish exposed to contaminated environments. 

39th Conference of the International Association for Great Lakes Research. Erindale College, 

Mississauga, ON. May 25-30. Invited Paper. 

Palace, VP, JF Klaverkamp. 1996. Oxidative stress in freshwater fish exposed to 

organochlorines. University of Manitoba Zoology Seminar Series, Winnipeg, Manitoba, March 

21. 

Brown, SB, VP Palace, JF Klaverkamp, J Fitzsimmons. 1995. Vitamins and oxidative stress in 

lake trout (Salvelinus namaycush) displaying early mortality syndrome. 2nd Workshop on 

Reproductive Disturbances in Fish. Stockholm, Sweden, November 20-23. 

Palace, VP, JF Klaverkamp, CL Baron, SB Brown. 1995. Retinoid metabolism in lake trout 

(Salvelinus namaycush) exposed to 3,3',4,4',5-pentachlorobiphenyl (PCB 126). 16th Annual 

Society of Environmental Toxicology and Chemistry Conference, Vancouver, B.C., Nov. 4-9. 

Palace, VP, TA Dick, SB Brown, CL Baron, JF Klaverkamp. 1994. Oxidative Stress in Lake 

Sturgeon Orally Exposed to 2,3,7,8-tetrachlorodobenzofuran. 15th Annual Society of 

Environmental Toxicology and Chemistry Conference, Denver, Colorado, October 29- 

November 3. 

Palace, VP, JF Klaverkamp, AS Kane. 1993. Phase I and Phase II metabolism enzyme activites 

and oxidative stress in lake trout from Lake Ontario. 14th Annual Society of Environmental 

Toxicology and Chemistry Conference, Houston, Texas, November 14-18. 

Brown, SB, VP Palace, DA Metner, JF Klaverkamp, WL Lockhart. 1993. Effect of 3,3',4,4',5-

pentachlorobiphenyl on retinoids, tocopherol and hepatic enzyme levels in lake trout (Salvelinus 

namaycush). 36th Conference of the International Association For Great Lakes Research, De 

Pere, Wisconsin, June 4-10. 

Palace, VP, JF Klaverkamp, SB Brown, WL Lockhart, DA Metner. 1992. An assessment of 

lipid peroxidative indices in lake char (Salvelinus namaycush) exposed to mixed-function 

oxidase inducing 2,3',4,4',5-pentachlorobiphenyl. 13th Annual Society of Environmental 

Toxicology and Chemistry Conference, Cincinnati, Ohio, November 8-12. 

Palace, VP, HS Majewski, JF Klaverkamp. 1990. Vitamin E and C deficiency and Cd toxicity in 

rainbow trout (Onchorhynchus mykiss). 19th Annual Fish Feed and Nutrition Workshop, 

Winnipeg, MB, October 15-19. 

Brown, SB, DF Malley, CL Baron, PSS Chang, E DeBruyn, R Evans, D Findlay, S Harrison, M 

Holoka, RV Hunt, J Klaverkamp, D Majewski, R McNichol, V Palace, CRanson, E Scherer, E 

Schindler, L Wesson. 1990. Cadmium whole lake experiment at the Experimental Lakes Area, 
29

northwestern Ontario and associated studies on cadmium. Proceedings of the 5th Annual 

LRTAP Workshop, St. John's Newfoundland, October 9-11.

Students Supervised Directly 

A) Postdoctoral Fellows 

 

 

C) Undergraduate Students 
30

Awards and Distinctions: 

2011 Sesion chair, SETAC Europe Milan Italy May 15-19 2011. 

2007 Editorial Board member of Environmental Toxicology and Chemistry 

2007 Session Chair for Aquatic Toxicity Workshop, Halifax NS. 

2006 Session Chair for ANCAP / SETAC Africa International Conference on Pesticide Use in 

 Developing Countries, Tanzania 

2001 Session Chair for Annual Society of Environmental Toxicology and Chemistry Meeting, 

Baltimore, MD 

2001 Session Chair for Aquatic Toxicology Workshop, Winnipeg, MB 

1999 Arnold Naimark Young Investigator Award for excellence in postdoctoral research by a 

Canadian in the field of Cardiovascular Research 

1996-98 Manitoba Health Research Council (MHRC) postdoctoral fellowship. 

1996 University of Manitoba postdoctoral fellowship. 

1994 NSERC/DFO Science Subvention Grant (Awarded to Dr. J.F. Klaverkamp on the basis of 

my written Ph.D. proposal) 

1993 Freshwater Institute graduate studentship