This document discusses the implications of inbreeding depression for laboratory ecotoxicology studies. It notes that captive test organisms used in these studies often experience reduced genetic diversity and inbreeding over time. Inbreeding depression can negatively impact organism fitness through expression of deleterious recessive alleles and loss of heterozygote advantage. This acts as an additional stressor on organisms, similar to pollutants and other stressors that laboratory studies aim to control for. Inbreeding depression can sensitize organisms to chemical toxicity, leading to overestimation of ecological risks. It can also potentially mask chemical effects if both stressors independently induce the same response. The document advocates for minimizing inbreeding in laboratory cultures to reduce these complications for ecotoxicology studies.

1. results that may defy current theory yet are consistent and
defensible as science.
Acknowledgment—Wayne Landis, Bob Lackey, and Ron
McCormick kindly suggested edits to my initial draft.
REFERENCES
[CMDSW] Center for Media and Democracy: Source Watch. 2015. The
advancement of sound science coalition. [cited 2015 October 21]. Available
from: http://www.sourcewatch.org/index.php?title=The_Advancement_of_
Sound_Science_Coalition
Greenpeace. 2013. Dealing in doubt. Introduction. [cited 2015 December 27].
Available from: http://climateandcapitalism.com/2013/09/10/climate-denial-
machine-vs-climate-science/
Lackey RT. 2007. Science, scientists, and policy advocacy. Conserv Biol 21:12–17.
McGarity TO. 2003. Our science is sound science and their science is junk science:
Science-based strategies for avoiding accountability and responsibility for
risk-producing products and activities. Kansas Law Rev 52:897–937.
McGarity TO, Wagner WE. 2008. Bending science: How special interests corrupt
public health research. Cambridge (MA): Harvard Univ Press. p 60–228.
Overton WR. 1982. Judgement in McLean versus the Arkansas Board of Education.
[cited 2016 January 5]. Available from: http://www.talkorigins.org/faqs/
mclean-v-Arkansas.html
Ravindran S. 2012. Barbara McClintock and the discovery of jumping genes. Proc
Natl Acad Sci USA 109:20198–20199.
[SETAC] Society of Environmental Toxicology and Chemistry. 1999. Technical issue
paper: Sound science. Pensacola (FL): SETAC. [cited 2015 October 21].
Available from: http://www.setac.org/default.asp?page=SETACTechPapers
THE PRACTICAL QUANTITATION LIMIT:
IMPLICATIONS FOR REGULATING SELENIUM IN THE
CONTEXT OF APPLYING AQUATIC LIFE GUIDELINES
IN NORTH AMERICA
Guy Gilron*y and James Downiez
yBorealis Environmental Consulting, North Vancouver, British
Columbia, Canada
zJRD Consulting, Quadra Island, British Columbia, Canada
*ggilron@yahoo.com
DOI: 10.1002/ieam.1780
Selenium water quality guidelines in support of human
health and ecological protection are currently in a state of
flux because of a growing body of literature on the toxicology
of Se in aquatic ecosystems (Janz et al. 2010). Much of this
recent work relates to effects on aquatic biota, in particular
egg-laying vertebrates (fish, birds, amphibians), which are
generally more sensitive to Se than are humans. Table 1
summarizes recent regulatory changes to aquatic life guide-
lines in North America; note the magnitude and range of
these guidelines (i.e., 1–5 mg/L).
Despite a growing understanding that tissue concentra-
tions of egg-laying vertebrates are the most appropriate
indicators of biological effects (Janz et al. 2010), there is still
a need to regulate Se based on the analysis and benchmarking
of aqueous concentrations, given logistical and economic
considerations for environmental monitoring programs.
Using the well-established adage “you can’t manage what
you don’t measure,” a key question that emerges when one
considers the regulation of Se in the context of water quality
is: “Can the existing analytical laboratory community
precisely and accurately measure Se down to the concen-
trations necessary to compare to aquatic life guidelines?” To
assess this issue, it is critical to understand how detection
limits are derived and defined.
The method detection limit (MDL) is a measure of method
sensitivity, within a given laboratory (precision); MDLs are
operator-, method-, laboratory-, and matrix-specific. The
MDL is defined in 40 FR Part 136 (Appendix B; http://www.
ecfr.gov/cgi-bin/text-idx?SID=30d649950843dcf4b8848f7-
ca9f35a3dmc=truenode=ap40.23.136_17.brgn=div9)
as: “the minimum concentration of a substance that can be
reported with 99% confidence that the analyte concentration
is greater than zero.” Because of normal day-to-day and run-
to-run analytical variability, MDLs may not always be
reproducible among 2 or more laboratories.
The practical quantitation limit (PQL) is a measure of a
laboratory capacities sensitivity (in other words, accuracy).
The PQL is defined as “the lowest achievable level of
analytical quantitation during routine laboratory operating
conditions within specified limits of precision and accu-
racy” (50 FR 46902, 1985). The US Environmental
Protection Agency (USEPA) uses the PQL to estimate or
evaluate the minimum concentration at which most
laboratories can be expected to reliably measure a specific
chemical parameter during day-to-day analyses.
Table 1. Summary of recent regulatory changes to aquatic life guidelines in North America
Jurisdiction
Guideline and/or
criterion (in mg/La
) Reference Notes on changes
United States (federal) Lotic: 3.1 USEPA 2015 (still in draft) Decreased from previous interim guideline
of 5 mg/L; now distinguishes between
lentic and lotic systems
Lentic: 1.2
Canada (national) 1 CCREM 1987 No change pending; derived using an
outdated approach
Province of British
Columbia (Canada)
2 Beatty and Russo 2014 The guideline document has been
updated but still has the same value
State of Kentucky (USA) 5 Kentucky Energy and
Environment Cabinet 2013
Applied using a tiered approach with
fish tissue concentration
a
Dissolved (i.e., that which can pass through a 0.45 mm filter).
Integr Environ Assess Manag 12, 2016—PM Chapman, Editor 593

2. The PQL has been used as a means of integrating
information on the performance of USEPA-approved analyti-
cal methods into the development of US drinking water
regulations (52 FR 25690; http://tftptf.com/CLW_Docs/
CLW1586A.pdf). Even though these studies were conducted
as part of the drinking water regulations, the analytical findings
are equally appropriate for the assessment of water quality for
the protection of aquatic life. Two studies pertaining to the
analyses of Se in drinking water (USEPA 2003, 2009) have
compared the PQL to the current maximum contaminant level
(MCL) and maximum contaminant level goal (MCLG) to
assess whether the PQL: 1) is adequate for measuring for MCL
and MCLG; or, 2) can it be adjusted downward to account
for improvements in analytical techniques and better inter-
laboratory consistency. The PQL is set for the concentration at
which 75% of laboratories are predicted to meet “acceptance
criteria.” The PQL can be derived in 1 of 2 ways, specifically:
the “MDL multiplier method,” a regulatory-derived MDL
multiplied by a factor of 5 or 10; or, using proficiency testing
(PT) study data (the preferred approach). The USEPA
Analytical Feasibility Studies (USEPA 2003, 2009) used the
PT study approach. The results from both Six-Year studies
indicated that the PQL for Se is 10 mg/L (or 0.01 mg/L)
(Figure 1), based on the fact that all of the passing rates and
both regression lines are well above the 75% threshold
(Figure 1).
Given the large scale of the 2 Six-Year studies (i.e., 80
laboratories participating in both studies), and the consistency
in the results reported in these studies, there are significant
implications to the PQL being assessed at a concentration of
10 mg/L, in light of above-mentioned changes to regulatory
aquatic life guidelines for Se (Table 1). First, background
concentrations of Se in natural waters (i.e., 0.5 mg/L), and
aquatic life guidelines (i.e., 1–5 mg/L) (Table 1) are generally in
the range of MDLs reported by most laboratories (i.e., 2 mg/L)
(USEPA 2009). Second, variability in analytical precision
increases near the MDL, which already introduces uncertainty
in the interpretation of the analyses; precision can be increased
with the use of higher-resolution analytical techniques, as the
MDL is method-specific. Finally, superimposed on this
“precision” uncertainty, the results of the 2 Six-Year studies
(USEPA 2003, 2009)—the “accuracy” component—indicate
that the aquatic life guidelines that we use for assessing
whether Se concentrations are safe are 50% to 88% lower than
the PQL, “the minimum concentration at which most
laboratories can be expected to reliably measure a specific
chemical parameter during day-to-day analyses.” In other
words, if we do not have confidence that we (collectively) can
accurately measure Se below an aqueous concentration of
10 mg/L, it is difficult to justify the use of benchmark
concentrations in the range of 1 to 5 mg/L, such as the current
aquatic life guidelines being recommended in North America.
To reduce these uncertainties, in addition to providing
benchmarks for the protection of aquatic life, regulatory
agencies in the United States (i.e., USEPA, state agencies) and
Canada (i.e., Environment and Climate Change Canada,
provincial environmental agencies) need to specify analytical
methods and/or techniques that will yield the precision and
Figure 1. Evaluation of Six-Year 1/ERA PT and Six-Year 2/ERA PT data—Se (after USEPA 2009).
594 Integr Environ Assess Manag 12, 2016—PM Chapman, Editor

3. accuracy required to ensure that aquatic life guidelines for Se
are applied appropriately.
REFERENCES
Beatty JM, Russo GA. 2014. Ambient water quality guidelines for selenium
technical report update. Victoria (BC): BC Ministry of Environment. 254 pp +
appendices.
[CCREM] Canadian Council of Resource and Environment Ministers. 1987.
Canadian water quality guidelines. Ottawa (ON): Task Force on Water Quality
Guidelines.
50 FR 46902. 1985. National primary drinking water regulations: Volatile synthetic
organic chemicals. Federal Register 50:46902–46906.
Janz D, Brooks ML, Chapman PM, DeForest D, Gilron G, Hoff D, Hopkins B,
McIntyre D, Mebane C, Palace V, et al. 2010. Selenium toxicity to aquatic
organisms. In: Chapman PM, Adams WJ, Brooks ML, Delos CG, Luoma SN,
Maher WA, Ohlendorf HM, Presser TS, Shaw DP, editors. Ecological assessment
ofseleniuminthe aquatic environment. Pensacola(FL): SETACPress. p 141–231.
Kentucky Energy and Environment Cabinet. 2013. Update to Kentucky water
quality standards for protection of aquatic life: acute selenium criterion and
tissue-based selenium chronic criteria. Frankfort (KY): Kentucky Energy and
Environment Cabinet. 35 pp + appendices.
[USEPA] US Environmental Protection Agency. 2003. Analytical feasibility support
document for the six-year review of national primary drinking water
regulations. Washington (DC): USEPA. EPA-815-R-03-003.
[USEPA] US Environmental Protection Agency. 2009. Analytical feasibility support
document for the second six-year review of existing national primary drinking
water regulations. Washington (DC): USEPA. EPA 815-B-09-003.
[USEPA] US Environmental Protection Agency. 2015. Draft aquatic life ambient
water quality criterion for Selenium—Freshwater 2015. Washington (DC):
USEPA. EPA 822-P-15-001.
INBREEDING DEPRESSION AS A COMPROMISING
FACTOR IN ECOTOXICOLOGICAL ASSAYS
Bryant S Gagliardi,*y Ary A Hoffmann,y and Vincent J Pettigrovey
yCentre for Aquatic Pollution Identification and Management,
School of Biosciences and Bio21 Institute, University of
Melbourne, Parkville, Victoria, Australia
*bgagliardi@student.unimelb.edu.au
DOI: 10.1002/ieam.1766
Laboratory ecotoxicological assays investigate causal relation-
ships between toxicants and biological stress responses. In field
experiments, it can be difficult to differentiate “true” pollutant
effects from thoseinduced by extraneous stressors (e.g., thermal
extremes, diseases, habitat degradation) (Chapman 1995).
These extraneous stressors can covary with pollution, are
sometimes undetectable, and may have complex interactions
with pollutants involving additivity, synergy, or antagonism.
Adding to this difficulty is the fact that most endpoints are
general stress—rather than pollution-specific—indicators.
Laboratory assays, by eliminating extraneous stressors, provide
an important tool for investigating cause–effect relationships.
Organisms for assays are usually reared in “in-house”
cultures. These cultures represent captive colonies. Captive
colonies are often subject to a reduction in genetic diversity
(relative to wild populations). This has been shown to be the
case for ecotoxicological cultures; genetic impoverishment is
common in cultures of the test aquatic insect Chironomus
riparius (Nowak, Vogt, et al. 2007) and fish Danio rerio (Coe
et al. 2009). Inbreeding, defined as reproduction between
closely related individuals, is a common consequence of
genetic impoverishment. It often results in inbreeding
depression (ID), which occurs when rare deleterious recessive
alleles become expressed when they occur as homozygotes as a
consequence of inbreeding; and through the loss of heterozy-
gote advantage. Inbreeding depression can result in a loss of
organism fitness.
Inbreeding depression therefore acts similarly to pollutants,
thermal extremes, diseases, and habitat degradation in that it
induces deleterious effects. It is potentially compromising to
the aims of laboratory ecotoxicology because it is a stressor that
is enhanced, rather than eliminated, in a laboratory setting.
Although ID has been studied extensively in conservation
biology, it has received relatively less attention in ecotoxico-
logy (Brown et al. 2009). Research suggests, however, that ID
is of concern. For some stress endpoints, ID interacts with
chemical toxicity, such that inbred organisms can have a
greater chemical sensitivity (and hence endpoint response)
than outbred animals (Nowak, Jost, et al. 2007; Brown et al.
2009). For other endpoints, ID has been shown to induce
certain endpoint responses independently of a chemical
response (Nowak, Jost, et al. 2007).
Some ramifications of ID for laboratory ecotoxicology have
been previously discussed by Brown et al. (2009). This
discussion focused on the interactive effects of ID and chemical
toxicity and on their potential to compromise ecological risk
assessments (ERAs). The authors point out that use of ID-
sensitized organisms in ERAs—which aim to extrapolate
laboratory findings to protect natural ecosystems—may lead to
overly protective/conservative ecological toxicity estimations.
They suggest strategies for minimizing inbreeding in cultures,
such as the use of large (1000) reproducing population sizes
or paired breeding programs. They also point out the
importance of quantifying the relationship between genetic
diversity levels and endpoint stress levels, and of identifying the
“critical” inbreeding levels above which fitness is compromised
(possibly the critical inbreeding coefficient of F ¼ 0.33) (Brown
et al. 2009).
In addition to sensitizing organisms (Nowak, Jost, et al.
2007; Brown et al. 2009), ID may also “mask” chemical effects.
This masking could occur when both ID and chemical
exposure independently (without interaction) induce a
particular endpoint response. Masking is illustrated in a
hypothetical scenario as follows. The increase in stress response
apparently associated with Chemical X in Figure 1A is
relatively small (the stress level at 40 mg/L being only 1.1Â
that of the control treatment), suggesting a low toxicity.
Figure 1B, however, differentiates ID-induced endpoint
responses from chemically induced endpoint responses for
this experiment and reveals the observed effects to be mostly
due to ID. Another experiment, exposing outbred organisms
to the same Chemical X concentrations (Figure 1C), induces
the same chemical effects. However, the absence of inbreeding
effects in this experiment means that the observed stress
response is much greater (10Â that of the control treatment at
40 mg/L), revealing a much greater toxicity. Masking in ERAs
may result in an underestimation of, or failure to detect,
chemical toxicity, producing insufficiently protective toxicity
values that risk ecosystem health.
Both sensitization (for endpoints subject to an ID-
chemical interaction) and masking (for ID-affected end-
points not subject to an interaction) may be problematic for
ERAs. Some standard assays require a sufficiently high fitness
in control organisms for a test to be considered valid (such as
a minimum control survival frequency in a mortality assay).
Integr Environ Assess Manag 12, 2016—PM Chapman, Editor 595