होम Environmental Toxicology and Chemistry Sources and fate of chiral organochlorine pesticides in western U.S. National Park ecosystems

Sources and fate of chiral organochlorine pesticides in western U.S. National Park ecosystems

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खंड:
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2011
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english
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DOI:
10.1002/etc.538
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Environmental Toxicology and Chemistry, Vol. 30, No. 7, pp. 1533–1538, 2011
# 2011 SETAC
Printed in the USA
DOI: 10.1002/etc.538

SOURCES AND FATE OF CHIRAL ORGANOCHLORINE PESTICIDES IN
WESTERN U.S. NATIONAL PARK ECOSYSTEMS
SUSAN A. GENUALDI,y KIMBERLY J. HAGEMAN,z LUKE K. ACKERMAN,§ SASCHA USENKO,k and
STACI L. MASSEY SIMONICH*y#
yDepartment of Chemistry, Oregon State University, Corvallis, Oregon, USA
zDepartment of Chemistry, University of Otago, Dunedin, New Zealand
§U.S. Food and Drug Administration, Center for Food Safety and Applied Nutrition, College Park, Maryland
kDepartment of Environmental Science, Baylor University, Waco, Texas, USA
#Department of Environmental and Molecular Toxicology, Oregon State University, Corvallis, Oregon, USA
(Submitted 23 December 2010; Returned for Revision 16 February 2011; Accepted 7 March 2011)
Abstract— The enantiomer fractions (EFs) of alpha-hexachlorocyclohexane (a-HCH), cis-, trans-, and oxychlordane, and heptachlor

epoxide were measured in 73 snow, fish, and sediment samples collected from remote lake catchments, over a wide range of latitudes, in
seven western U.S. National Parks/Preserves to investigate their sources, fate, accumulation and biotransformation in these ecosystems.
The present study is novel because these lakes had no inflow or outflow, and the measurement of chiral organochlorine pesticides (OCPs)
EFs in snowpack from these lake catchments provided a better understanding of the OCP sources in the western United States, whereas
their measurement in fish and sediment provided a better understanding of their biotic transformations within the lake catchments.
Nonracemic a-HCH was measured in seasonal snowpack collected from continental U.S. National Parks, and racemic a-HCH was
measured in seasonal snowpack collected from the Alaskan parks, suggesting the influence of regional sources to the continental U.S.
parks and long-range sources to the Alaskan parks. The a-HCH EFs measured in trout collected from the lake catchments were similar to
the a-HCH E; Fs measured in seasonal snowpack collected from the same lake catchments, suggesting that these fish did not biotransform
a-HCH enantioselectively. Racemic cis-chlordane was measured in seasonal snowpack and sediment collected from Sequoia, indicating
that it had not undergone significant enantioselective biotransformation in urban soils since its use as a termiticide in the surrounding
urban areas. However, nonracemic cis-chlordane was measured in seasonal snowpack and sediments from the Rocky Mountains,
suggesting that cis-chlordane does undergo enantioselective biotransformation in agricultural soils. The trout from these lakes showed
preferential biotransformation of the (þ)-enantiomer of cis-chlordane and the ()-enantiomer of trans-chlordane. Environ. Toxicol.
Chem. 2011;30:1533–1538. # 2011 SETAC
Keywords—Chiral

Organochlorine pesticides

Enantiomer fraction

Chlordane

Hexachlorocyclohexane

The enantiomer fractions (EFs) of chiral OCPs have been
used to identify their sources in ecosystems, as well as their
accumulation and biotransformation in food webs [12–14].
Chiral pesticides are composed of two enantiomers and are
typically manufactured in racemic (50:50) form. The enantiomers have the same physical and chemical properties and are
affected by abiotic processes in the same way [15]. However,
biotic processes, involving enzymatic activity, may result in
enantioselective biotransformation [15]. Recent use of the
pesticide or nonenantioselective biotransformation will result
in a racemic EF, whereas enantioselective biotransformation
will result in a nonracemic EF. As a result, EFs have been used
to apportion chiral organic pollutant sources to water bodies
[12], identify their emission source regions to the atmosphere
[15], and track their fate and biotransformation in food webs
[13].
The present study is novel because the WACAP lake catchments span a broad latitude range (36–68 decimal degrees N),
from southern California to the northern point of Alaska. In
addition, these lakes have no inflow or outflow through rivers or
streams. Therefore, the OCP sources to the lakes include only
atmospheric deposition processes, and their removal includes
abiotic or biotic transformation processes within the lake catchment. The measurement of chiral OCP EFs in snowpack from
these lake catchments provides a better understanding of the
OCP sources in the Western U.S., whereas their measurement in
fish and sediment provides a better understanding of the biotic

INTRODUCTION

Organochlorine pesticides (OCPs) are persistent in the environment and have been shown to undergo long-range atmospheric transport and accumulation in remote high-elevation and
high-latitude ecosystems [1–5]. Identifying OCP sources to
these sensitive ecosystems is important, because many of them
are toxic and can potentially bioaccumulate in food webs [1]. A
major objective of the Western Airborne Contaminants Assessment Project (WACAP, http://www2.nature.nps.gov/air/
Studies/air_toxics/wacap.htm) was to determine the sources
of semivolatile organic contaminants in high-elevation and
high-latitude ecosystems in Western U.S. National Parks
[3,6]. Organochlorine pesticides have been measured in seasonal snowpack, lake water, sediment, vegetation, and fish from
these ecosystems [1–4,7,8]. Snow is a major route of OCP
deposition to these ecosystems, because it is an effective
scavenger of both gas- and particulate-phase OCPs from the
atmosphere [2,7,9,10]. The melting of the seasonal snowpack
results in the release of OCPs to the aquatic or terrestrial
ecosystem and is a source of OCPs to the lake water, sediment,
and fish [11].
All Supplemental Data may be found in the online version of this article.
* To whom correspondence may be addressed
(staci.simonich@orst.edu).
Published online 1 April 2011 in Wiley Online Library
(wileyonlinelibrary.com).
1533

1534

Environ. Toxicol. Chem. 30, 2011

transformations within the lake catchments. In the present
study, the EFs of alpha-hexachlorocyclohexane (a-HCH),
cis-, trans-, and oxychlordane, and heptachlor epoxide B were
measured in 73 snowpack, fish, and sediment samples collected
from remote lake catchments in seven western U.S. National
Parks. These data were used to identify the sources of these
pesticides and assess their fate, accumulation, and biotransformation in these remote ecosystems.
MATERIALS AND METHODS

Sampling sites

Seasonal snowpack, sediment cores, and fish: brook trout
(Salvelinus fontinalis), lake trout (Salvelinus namaycush), and,
on occasion. cutthroat trout (Oncorhynchus clarki), rainbow
trout (Oncorhynchus mykiss), and burbot (Lota lota), were
collected from Emerald and Pear Lake catchments in Sequoia
National Park, Lone Pine, and Mills Lake catchments in Rocky
Mountain National Park, LP19, and Golden Lake catchments in
Mt. Rainier National Park, Oldman and Snyder Lake catchments in Glacier National Park, Wonder Lake catchment in
Denali National Park, Burial Lake catchment in Noatak
National Preserve, and Matcharak Lake catchment in Gates
of the Arctic National Park and Preserve between 2003 and
2005 as part of the U.S. National Park Service’s WACAP
(Supplemental Fig. S1, Supplemental Table S1) [1–4]. Seasonal
snowpack samples were collected from one to three sites at each
park at the end of the snow accumulation season in 2003, 2004,
and 2005 [2,7]. Sediment cores were collected from the deepest
point in each lake, sectioned, and dated using 210Pb and
137
Cs [4], and 10 fish per lake were collected and analyzed
[1]. Site and sampling procedure details have been previously
reported [1–4].
Analysis of chiral OCPs

The analytical methods for the extraction, cleanup, and
instrumental analysis of OCPs in snow, fish, and sediment have
been reported previously, along with the chemicals, standards,
and solvents used [1–4,16]. The OCP concentrations in each of
these matrices were determined by using isotope dilution gas
chromatographic mass spectrometry and are reported elsewhere
[1–4].
Enantioselective analysis of the OCPs and OCP degradation
products was performed using gas chromatographic mass spectrometry with electron capture negative ionization and a DB-5
column (28 m, 0.25 mm inner diameter, 0.25-mm film thickness,
J&W Scientific) connected to a BGB Analytik chiral column
(10 m, 0.25 mm inner diameter, 0.25-mm film thickness, BGB
Analytik, Germany). The details of the gas chromatography
temperature program, instrumental parameters, and ions monitored have been reported elsewhere [17]. The BGB column was
used to separate the enantiomers of a-HCH, oxychlordane,
trans-chlordane, cis-chlordane, heptachlor epoxide and
o,p0 -DDT. All of these chiral OCPs were measured above
the detection limit for enantioselective analysis except for
o,p0 -DDT. Pure enantiomer standards were used to determine
the elution order of a-HCH, cis-chlordane, and trans-chlordane
(Dr. Ehrenstorfer). The enantiomer fractions (EFs) of chiral
OCPs were calculated using the following equation [18]:
EF ¼

area of the ðþÞ enantiomer
area of the ðþÞ enantiomer þ area of the ðÞ enantiomer
(1)

S.A. Genualdi et al.

An EF of 0.5 indicates a racemic mixture. Pure enantiomer
standards were not available for heptachlor epoxide and
oxychlordane, and the elution order of the enantiomers could
not be determined. In these cases, the EFs were calculated using
the following formula:
EF ¼

area of the ð1Þ eluting enantiomer
area of the ð1Þ eluting enantiomer þ area of the
ð2Þ eluting enantiomer

(2)

Manual integration of peak area was conducted using macro
smoothed chromatograms in MSD ChemStation (G1701DA).
The racemic ranges of the chiral OCPs were calculated from
seven replicate injections of the racemic standard (25 pg/mL).
The 95% confidence intervals for the racemic ranges were
0.499  0.0095 for a-HCH, 0.499  0.0052 for trans-chlordane, and 0.497  0.0059 for cis-chlordane. Because the EF
precision of the standards was four decimal places, all EFs are
reported to three decimal places. Ion ratios were monitored in
each sample to ensure no matrix interferences, and the ratios
were required to be within 20% of the ratios of the standards.
Because of matrix interferences with the (þ)-enantiomer, the
EF of trans-chlordane was not determined in sediment extracts.
As a result, we have limited our interpretation of the (þ)enantiomer of trans-chlordane to seasonal snowpack and fish.
Because of the relatively low OCP concentrations in the
samples and the higher detection limit for enantioselective
analysis, the measured OCP concentrations were used to prioritize the samples for enantioselective analysis. All samples
with OCP concentrations above the detection limit were analyzed [1,2,4,7,16]. Within a park, snowpack, sediment, and fish
samples with the highest OCP concentrations were analyzed
first. If the concentrations were above the detection limit,
additional samples from the park were analyzed. In all, 25
snowpack, 18 sediment sections, and 30 fish samples were
analyzed.
The detection limit for enantioselective analysis was defined
as a signal-to-noise ratio of 3:1 [19], and these values can be
found in Supplemental Table S2. For a-HCH, this was approximately 0.05 ng/L in snow and 0.09 ng/g in fish. The signal-tonoise ratio for a-HCH was less than 3:1 in all of the sediment
samples, resulting in a detection limit greater than 0.8 ng/g. For
trans-chlordane, the detection limits were approximately
0.01 ng/L for snow, 0.07 ng/g for sediment, and 0.04 ng/g for
fish. The detection limit for cis-chlordane was approximately
0.02 ng/g for fish, 0.2 ng/g for sediment, and 0.02 ng/L for snow.
Oxy-chlordane and heptachlor epoxide were measured above
the detection limit only in fish samples (detection limits of
approximately 0.02 ng/g and 0.08 ng/g, respectively). All laboratory and field blanks had concentrations below the detection
limit for enantioselective analysis.
RESULTS AND DISCUSSION

Sources of chiral pesticides

Technical HCH, which is considered the primary source of
a-HCH to the environment, was used in the United States until
1978, in Canada until 1971, and in Asia until 2000 [20,21]. With
an atmospheric half-life of 115 d and physical–chemical properties that result in long-range atmospheric transport, a-HCH
undergoes cold-condensation to remote regions ([22]; http://
www.atsdr.cdc.gov/ToxProfiles/tp43.pdf). The measurement of
nonracemic a-HCH in snowpack indicates that its source to the
lake catchment is revolatilization from prior uses after biotic

Enantiomer fractions used to study a remote ecosystem

transformation, whereas racemic a-HCH in snowpack indicates
a fresh source or a source that has only undergone abiotic
transformation. Within a park, the seasonal snowpack was
typically either consistently racemic or nonracemic for a given
OCP year to year (Supplemental Table S1[A]). However,
the seasonal snowpack EF values varied somewhat among
sites and years, within a park. For example, the seasonal
snowpack a-HCH EFs in Mt. Rainier ranged from 0.520 to
0.562, but were consistently nonracemic and depleted in the
()-enantiomer.
A t test revealed a significant difference ( p value < 0.001)
between the mean racemic a-HCH EFs (0.505  0.0057) measured in seasonal snowpack in Alaska (Noatak, Gates of the
Arctic, and Denali) and the mean nonracemic a-HCH EFs
(0.529  0.0131) measured in seasonal snowpack in all other
parks (Sequoia, Glacier, Mt. Rainier, and Rocky Mountain)
(Fig. 1A, Supplemental Table S1A). The seasonal snowpack
historic-use pesticide concentrations at WACAP sites in the
continental United States (Sequoia, Olympic, Mt. Rainier,
Glacier, and Rocky) result from their volatilization from
regional agricultural soils, followed by regional atmospheric
transport and deposition [2,7]. However, the seasonal snowpack
historic-use pesticide concentrations in the Alaska parks
(Denali and Gates of the Arctic/Noatak) result from long-range
atmospheric transport and deposition [2,7]. No statistically
significant correlations were observed between the OCP concentrations and EF values.
The seasonal snowpack a-HCH EFs were negatively correlated with site latitude (r2 ¼ 0.396, p ¼ 0.007) and positively
correlated with mean site winter temperature (r2 ¼ 0.455,
p ¼ 0.003) (Supplemental Fig. S2). A positive correlation
was also observed with site elevation (Supplemental Fig.
S2). The correlations with site latitude and mean winter temperature confirm the observation that a-HCH EFs in seasonal
snowpack were closer to racemic at higher-latitude, lowertemperature sites (the Alaskan parks) and that a-HCH undergoes cold condensation to the Alaskan parks. We measured
nonracemic a-HCH (enhanced in the (þ)-enantiomer) in seasonal snowpack at Sequoia, Glacier, Mt. Rainier, and Rocky
Mountain. Its likely source was volatilization from areas where
enantioselective biodegradation had occurred, such as regional
agricultural soils or the Pacific Ocean, followed by transport and
deposition to the site. Nonracemic a-HCH (also enhanced in the

Environ. Toxicol. Chem. 30, 2011

1535

[þ]-enantiomer) has been measured in air at a remote, marine
site on the Pacific coast of the United States and was attributed
to volatilization from background soils at the site or volatilization from the Pacific Ocean [17]. The racemic a-HCH in
seasonal snowpack in the Alaska parks may be explained by
long-range atmospheric transport. Racemic a-HCH has been
measured in other matrices in the Arctic [23], as well as in free
tropospheric air masses [17,23].
Technical chlordane was banned in the United States for
agricultural use in 1983 and from residential termiticide use in
1988. As a result, its presence in U.S. air is primarily from
volatilization from agricultural soils in rural areas (showing a
nonracemic EF) and soils surrounding the foundations of homes
in urban areas (showing a racemic EF) [24,25]. Racemic EFs
in urban soils have been previously attributed to the higher
chlordane concentrations in urban soils suppressing biotransformation [19]. Significant differences also have been measured
between nonracemic chlordane EFs in tilled-planted and
untilled soils, suggesting that tilling soils and growing crops
can produce enantioselective changes [19]. We measured racemic cis- and trans-chlordane in seasonal snowpack from
Sequoia and nonracemic cis- and trans- chlordane in Glacier,
Noatak, and Gates of the Arctic (Fig. 1B, C; Supplemental
Table S1[A]). The seasonal snowpack chlordane EFs were not
correlated with site latitude, mean winter temperature, or
elevation (Supplemental Fig. S2), indicating that chlordane
did not undergo cold condensation to the Alaskan parks to
the degree that a-HCH did. The Rocky and Mt. Rainier seasonal
snowpack chlordane concentrations were below the detection
limit for enantioselective analysis.
Of the WACAP parks, Sequoia has the highest population
density within 300 km of the park (12,841,589 people) [3]. As
part of a global passive air sampling study, some of the highest
chlordane concentrations in the world were measured near Los
Angeles, CA, a site approximately 250 km south of Sequoia
National Park [26,27]. Previous studies in the Western United
States have also measured racemic cis- and trans-chlordane EFs
in air masses influenced by California [17]. Because the chlordane concentration in urban soils is approximately 25 to 100
times higher than agricultural soils [19], re-volatilization of
racemic chlordane from urban soils may be more significant
than re-volatilization of nonracemic chlordane from agricultural
soils at parks surrounded by high population densities.

Fig. 1. Enantiomer fractions (EFs) of (A) alpha-hexachlorocyclohexane (a-HCH), (B) trans-chlordane, and (C) cis-chlordane EFs in seasonal snowpack,
sediment, and fish from lake catchments in Sequoia (SEKI), Rocky Mountain (ROMO), Mt. Rainier (MORA), Glacier (GLAC), Denali (DENA), Gates of the Arctic
(GAAR), and Noatak (NOAT). The racemic range is shown by the solid lines on each graph. Matrices with no data points were below the detection limit.

1536

Environ. Toxicol. Chem. 30, 2011

In contrast, Glacier has a relatively low population density
but a relatively high cropland intensity (including the surrounding agricultural land in Canada) [3]. Because both cropland
intensity and population density were low surrounding the
Alaskan parks [2,3,6], the nonracemic chlordane measured in
Noatak and Gates of the Arctic seasonal snowpack was likely
attributable to long-range transport from agricultural soils in
Asia and North America. Together, our data suggest that the
chlordane EF in seasonal snowpack at WACAP sites is influenced by regional transport from urban soils (racemic in
Sequoia), regional transport from agricultural soils (nonracemic
in Glacier), and long-range transport from agricultural soils
(nonracemic in Noatak and Gates of the Arctic).
Fate, accumulation, and biotransformation of OCPs

Sediment. Only cis-chlordane was quantifiable in sediment
samples because of matrix interferences in the measurement
of trans-chlordane and a-HCH concentrations were below the
detection limit for enantioselective analysis. In Sequoia, racemic cis-chlordane was measured in seasonal snowpack and
sediment from both Emerald and Pear Lake catchments. This
indicated that the sediment cis-chlordane EF reflected the EF of
snow in both lake catchments (Fig. 1C and Supplemental Table
S1[A,B]). Previous studies measured racemic cis-chlordane in
Long Island Sound sediment and flooded soils, which suggested
that the anaerobic conditions prevented enantioselective biodegradation [28,29].
Racemic cis-chlordane was also measured in sediment from
Mt. Rainier (Fig. 1C and Supplemental Fig. S3). Of the national
parks studied, Mt. Rainier has the second highest population
density within 300 km of the park, at 8,703,126 people [3].
Although the cis-chlordane EF in the seasonal snowpack at Mt.
Rainier was below the detection limit for enantioselective
analysis, the sediment from Mt. Rainier suggests that, like
Sequoia, this park receives racemic chlordane from volatilization from surrounding urban soils.
A one-way analysis of variance indicated that the mean cischlordane EFs, measured in sediment from the different parks,
were not equal. In addition, Scheffe confidence intervals
showed that the nonracemic cis-chlordane EFs (depleted in
the []-enantiomer) in Mills Lake and Lone Pine Lake sedi-

S.A. Genualdi et al.

ment in Rocky Mountain were not significantly different from
each other (Supplemental Fig. S4). However, the nonracemic
cis-chlordane EFs in Rocky Mountain sediment were significantly different from the racemic cis-chlordane EFs in Sequoia
and Mt. Rainier sediment (Supplemental Fig. S4). A previous
study has shown that cis-chlordane is depleted in the ()enantiomer in agricultural soils from Alabama, the Midwestern
United States, Connecticut, Hawaii, and the United Kingdom
[30]. Our data suggest that the nonracemic cis-chlordane in
Rocky Mountain sediment is attributable to volatilization from
regional agricultural soils, whereas the racemic cis-chlordane in
Sequoia and Mt. Rainier sediments is attributable to volatilization from surrounding urban areas, where chlordane was commonly used as a termiticide.
Fish. Aquatic organisms are exposed to chiral pesticides
through the water column, diet, and sediment [12]. For hydrophobic pesticides, such as chlordane, with log octanol–water
partition coefficients greater than 5, pesticide uptake by fish
through the food web is more significant than uptake through
the water column [12]. In general, the EFs of chiral pesticides
increase in their deviation from racemic in the following order:
air < water < soil < biota because of increasing enantioselective biotransformation [13].
Alpha-hexachlorocyclohexane has a relatively low log octanol–water partition coefficient (3.8) [31] but is the most
abundant OCP in air and water in northern latitudes [32–35].
We measured a-HCH above the detection limit for enantioselective analysis in only nine of the 30 fish samples (six of the
nine were from Alaskan parks), and the EFs ranged from 0.309
to 0.507 (Fig. 2B and Supplemental Table S1[C]). The Wonder
Lake fish in Denali had a-HCH EFs significantly (two-sided
p < 0.0013) lower than all other fish in this study (0.309–0.348).
This may be attributable to increased enhanced biological
activity (higher specific conductivity and suspended solids)
in Wonder Lake compared with the other lakes [8]. Excluding
the Denali fish, the a-HCH EFs in fish reflected the a-HCH EFs
in seasonal snowpack from the same lake catchment (Fig. 1A).
Cis- and trans-chlordane were measured above the detection
limit for enantioselective analysis in 15 of the 30 fish samples
and 10 of the 30 fish samples, respectively. Nonracemic cischlordane EFs ranged from 0.043 to 0.466 and was measured in

Fig. 2. Enantiomer fractions (EFs) (A) Oxy-, cis-, and trans-chlordane and (B) alpha-hexachlorocyclohexane (a-HCH) and heptachlor epoxide EFs in fish
collected from Sequoia, Rocky Mountain, Mt. Rainier, Glacier, Denali, and Gates of the Arctic. The black line represents a racemic EF of 0.5. The racemic ranges
are different for each compound and are not displayed in the figure. Each symbol represents one measurement. [Color figure can be seen in the online version of this
article, available at wileyonlinelibrary.com]

Enantiomer fractions used to study a remote ecosystem

all parks, except Sequoia. Nonracemic trans-chlordane EFs
ranged from 0.582 to 0.900 and were measured in all parks,
except Glacier and Gates of the Arctic (Fig. 2A and Supplemental Table S1[C]). Cis-chlordane was consistently depleted
in the (þ)-enantiomer in fish, whereas trans-chlordane was
consistently depleted in the ()-enantiomer in fish (Fig. 2A,
Supplemental Fig. S1[C]). Previous measurements made in
rainbow trout and Arctic char found the same depletion patterns
[36,37]. The difference between the nonracemic chlordane EFs
in fish and the racemic chlordane EFs in seasonal snowpack
from the same lake catchments is likely attributable to enantioselective biotransformation in the fish or biotransformation
in the food web and bioaccumulation through trophic transfer
(Figure 1B, 1C).
The results of the present study suggest that cis- and transchlordane are enantioselectively biotransformed in these trout
species to a greater extent than a-HCH. A study by Wong et al.
[37] suggested that trout do not enantioselectively biotransform
a-HCH and that nonracemic residues of a-HCH in rainbow
trout are caused by uptake of nonracemic a-HCH through the
diet [37]. In this same study, trout were also fed racemic
chlordane and, after 13 d, the cis- and trans-chlordane EFs
were significantly nonracemic. These laboratory studies support
our hypothesis that fish from these national parks can enantioselectively biotransform cis- and trans-chlordane but not aHCH.
Oxy-chlordane (a metabolite of chlordane) was measured in
24 of the 30 fish samples and ranged in EF from 0.432 to 0.560
in fish from all parks (Fig. 2A). A previous study with trout fed
racemic chlordane measured oxy-chlordane a few days after
measuring trans-chlordane in the fish [37]. However, the EF of
oxy-chlordane was not measured [37]. Previous studies measured similar EF values of nonracemic oxy-chlordane in ringed
seal and polar bears, indicating bioaccumulation of oxy-chlordane through the food web [14]. Nonracemic heptachlor epoxide (a metabolite of heptachlor) was measured in 5 of the 30 fish
samples (average of 0.619  0.137) and was measured in fish
from the Alaskan parks and Mt. Rainier (Fig. 2B). Although
heptachlor epoxide was below the detection limit for enantioselective analysis in seasonal snowpack, previous studies have
measured its nonracemic EF in Lake Ontario air (0.65) and Lake
Superior (0.67) and North Pole (0.62) surface water [13]. A
previous study determined that fish are unable to biotransform
heptachlor epoxide [38]. This suggests that the nonracemic
heptachlor epoxide in WACAP fish may not be attributable
to enantioselective biotransformation, but may be caused by the
deposition and accumulation of nonracemic heptachlor epoxide
in the food web.
No significant correlations were found between the fish aHCH, oxy-chlordane, cis-chlordane, and trans-chlordane EFs
and fish age and weight. In addition, a t test showed no
significant difference between the a-HCH, oxy-chlordane,
cis-chlordane, and trans-chlordane EFs because of fish sex or
trout species. Previous studies have measured different EFs in
different fish species. However, in this present study, the
enantioselective biotransformation of cis- and trans-chlordane
in trout appeared to be independent of fish age, weight, sex, and
species. This may be an artifact of the relatively small sample
size in this study.
The measurement of OCP EFs in different parts of a highelevation or high-latitude lake catchment is useful for identifying sources and tracking the fate of OCPs in the lake catchment,
from deposition (snowpack) to accumulation (fish and sediment). In this study, the EFs showed that the OCP sources to

Environ. Toxicol. Chem. 30, 2011

1537

these western U.S. lake catchments were dependent on the
park’s latitude, distance from populated regions, and proximity
to agriculture. The EFs were also used to gain a better understanding of which OCPs underwent biotic degradation once
deposited to the lake catchment and which bioaccumulated in
the food web.
SUPPLEMENTAL DATA

Table S1. Sample information and a-HCH, trans-chlordane
(TC), and cis-chlordane (CC) EFs in seasonal snowpack, sediment, and fish at WACAP sites.
Table S2. Detection limits for enantioselective analysis
which corresponds to a signal:noise ratio of 3:1.
Fig. S1. Map of National Parks sampled.
Fig. S2. Correlations between the a-HCH and trans-chlordane EFs in seasonal snowpack with site temperature, latitude
and elevation.
Fig. S3. Cis-chlordane EFs in surficial and core sediment
collected from national parks.
Fig. S4. Scheffe confidence intervals of cis-chlordane EFs
measured in sediment at National Parks. (1.3 MB DOC)
Acknowledgement—This work is part of the Western Airborne Contaminants
Assessment Project. This publication was made possible in part by grant
P30ES00210 from the United States National Institute of Environmental
Health Sciences (NIEHS), part of the National Institute of Health (NIH). Its
contents are solely the responsibility of the authors and do not necessarily
represent the official view of the NIH, NIEHS. This work was partially funded
by the U.S. Environmental Protection Agency and the Department of the
Interior. It has been subjected to review by these government entities and
approved for publication. Approval does not signify that the contents reflect
the views of the U.S. government, nor does mention of trade names or
commercial products constitute endorsement or recommendation.
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