World-wide Measurements of Atmospheric CO 2 and Other Trace Species Using Commercial Airlines
EXPOSURE OF GAME SPECIES TO TRACE ELEMENTS AND …
Transcript of EXPOSURE OF GAME SPECIES TO TRACE ELEMENTS AND …
EXPOSURE OF GAME SPECIES TO TRACE ELEMENTS AND
RADIOCESIUM ON THE SAVANNAH RIVER SITE IN SOUTH CAROLINA
by
RICKI ELAINE OLDENKAMP
(Under the Direction of James C. Beasley)
ABSTRACT
Despite the widespread harvest and consumption of game by recreational hunters,
there are few data available regarding contaminant burdens in many commonly harvested
wildlife species. I sampled wild pigs (Sus scrofa), gray squirrels (Sciurus carolinensis),
and waterfowl from a contaminated Department of Energy site to quantify concentrations
of trace elements and radiocesium in muscle and liver tissues for assessment of potential
human health risks from the consumption of game, and contaminant accumulation rates
in tissues. Concentrations varied among collection locations and species, although
waterfowl collected from a coal ash basin consistently had high levels of trace element
burdens (especially Selenium) and those from areas with known radiological
contamination had elevated radiocesium concentrations, often exceeding limits
established by the European Economic Community for safe human consumption.
INDEX WORDS: Coal Combustion Waste, Game Meat, Gray Squirrels, Human Consumption Risk, Waterbirds, Waterfowl, Wild Pigs, Radiocesium, Trace Elements
EXPOSURE OF GAME SPECIES TO TRACE ELEMENTS AND
RADIOCESIUM ON THE SAVANNAH RIVER SITE IN SOUTH CAROLINA
by
RICKI ELAINE OLDENKAMP
B.S., Biology, Northern Michigan University, 2013
A Thesis Submitted to the Graduate Faculty of The University of Georgia in Partial
Fulfillment of the Requirements for the Degree
MASTER OF SCIENCE
ATHENS, GEORGIA
2016
EXPOSURE OF GAME SPECIES TO TRACE ELEMENTS AND
RADIOCESIUM ON THE SAVANNAH RIVER SITE IN SOUTH CAROLINA
by
RICKI ELAINE OLDENKAMP
Major Professor: James C. Beasley
Committee: Karl V. Miller William A. Hopkins Electronic Version Approved: Suzanne Barbour Dean of the Graduate School The University of Georgia May 2015
v
ACKNOWLEDGEMENTS
I have learned so much during the last two years, and am extremely thankful to
Jim Beasley for giving me a chance when I didn’t have a traditional wildlife background.
My first friend upon starting the program has been by my side throughout this journey
and also been my “bird dog” for duck research, Chris Leaphart I never would have gotten
through this without you. I am extremely thankful for my friends, Becky, Betsy, and
Nathan, y’all have been very supportive and taught me things from hunting to Cajun
cooking and I will forever cherish those memories. I am also thankful to Larry Bryan and
Bobby Kennamer for their support and advice. Also to Bill Hopkins, whose brain I envy
and who is super kind to share his smarts with me. To my favorite professor, Karl Miller,
you are funny and so damn smart, I loved your habitat class and that course made me feel
I had truly found the right path for my career, I hope I make you proud. I am also grateful
for the SREL folks, there have been many fun times and trying times out there in the
middle of nowhere South Carolina, glad we had each other for both. If I did not have so
much love and support from my wonderful father, Rick Oldenkamp, and sister, Amanda
Szabo, I most definitely would have given up in some of my most stressful moments, I
love you both more than I can explain. Also to my favorite person in the world, my
brother TJ Oldenkamp, you have inspired me greatly with all that you have accomplished
over these last few years, I am so proud of the man you are and it has made me want to
meet you in that place of excellence, I am extremely glad we are friends and your no-
nonsense advice propelled me forward to the finish, thank you.
vi
TABLE OF CONTENTS
Page
ACKNOWLEDGEMENTS .................................................................................................v
LIST OF TABLES ........................................................................................................... viii
LIST OF FIGURES ......................................................................................................... xiii
CHAPTER
1 INTRODUCTION AND LITERATURE REVIEW .........................................1
LITERATURE CITED ................................................................................7
2 TRACE ELEMENTS AND RADIOCESIUM IN GAME SPECIES AND
HUMAN CONSUMPTION RISKS ................................................................18
ABSTRACT ...............................................................................................19
INTRODUCTION .....................................................................................20
METHODS ................................................................................................24
RESULTS ..................................................................................................35
DISCUSSION ............................................................................................41
MANAGEMENT IMPLICATIONS .........................................................47
LITERATURE CITED ..............................................................................67
3 WATERFOWL EXPOSURE TO COAL COMBUSTION WASTES AND
HUMAN CONSUMPTION RISKS ................................................................85
ABSTRACT ...............................................................................................86
INTRODUCTION .....................................................................................87
v
METHODS ................................................................................................91
RESULTS ..................................................................................................99
DISCUSSION ..........................................................................................101
MANAGEMENT IMPLICATIONS .......................................................104
LITERATURE CITED ............................................................................116
4 RADIOCESIUM IN WATERFOWL/WATERBIRDS FROM A RETIRED
NUCLEAR REACTOR COOLING RESERVOIR .......................................129
ABSTRACT .............................................................................................130
INTRODUCTION ...................................................................................131
METHODS ..............................................................................................134
RESULTS ................................................................................................141
DISCUSSION ..........................................................................................143
CONCLUSION ........................................................................................146
LITERATURE CITED ............................................................................154
5 CONCLUSION ..............................................................................................160
vi
LIST OF TABLES
Page
Table 2-1: Sample sizes of game species collected from the Savannah River Site (SRS)
and analyzed for trace elements and radiocesium in 2012-2015; sex ratios are also
indicated. ................................................................................................................48
Table 2-2: Comparisons of trace element concentrations (ppm, dry mass) in muscle and
liver tissues of wild pigs collected from the Savannah River Site (SRS) and those
collected from five counties in Georgia (GA) in 2012-2015. See Table 1 for
sample sizes of muscle and liver tissues analyzed. ................................................49
Table 2-3: Correlations among trace element concentrations in muscle (above diagonal)
and liver (below diagonal) tissues of wild pigs collected from the Savannah River
Site (SRS) in 2012-2015 (n=88 muscle, n=30 liver). Only those elements with
more than 50% of the values above detectable limits are included. Correlations
between muscle and liver samples for individual elements are presented on the
diagonal in bold. .....................................................................................................50
Table 2-4: Correlations among trace element concentrations in muscle (above diagonal)
and liver (below diagonal) tissues of wild pigs collected from five counties in
Georgia (GA) in 2012-2015 (n=20 muscle, n=20 liver). Only those elements with
more than 50% of the values above detectable limits are included. Correlations
between muscle and liver samples for individual elements are presented on the
diagonal in bold. .....................................................................................................51
vii
Table 2-5: Descriptive statistics for radiocesium concentrations in muscle and liver
tissues of wild pigs collected from the Savannah River Site (SRS) and from five
counties in Georgia (GA) in 2012-2015. ...............................................................52
Table 2-6: Radiocesium concentrations (Bq/g, dry mass) in muscle and liver tissues of
wild pigs collected from the Savannah River Site (SRS) and from five counties in
Georgia (GA) from 2012-2015. .............................................................................53
Table 2-7: Comparisons of trace element concentrations (ppm, dry mass) in muscle of
squirrels collected near the D-Area ash basins and squirrels collected from all
other locations on the Savannah River Site (SRS) in 2012-2015. .........................54
Table 2-8: Descriptive statistics for radiocesium concentrations in muscle tissue of
squirrels collected from the Savannah River Site (SRS) in 2012-2015. ................55
Table 2-9: Comparison of radiocesium concentrations (Bq/g, dry mass) in muscle tissue
of squirrels collected from four different locations on the SRS in 2012-2015.a ....56
Table 2-10: Waterfowl and waterbird species collected on the Savannah River Site (SRS)
in 2012-2015, with the scientific names, alpha codes, guild groupings, and sample
sizes in various analyses. .......................................................................................57
Table 2-11: Comparisons of trace element concentrations (ppm, dry mass) in muscle and
liver tissues of diving ducks collected from D-Area ash basins and diving ducks
collected from other water bodies on the Savannah River Site (SRS) in 2012-
2015. ....................................................................................................................58
Table 2-12: Concentrations of trace elements (ppm, dry mass) in muscle and liver tissues
of dabbling ducks collected from the Savannah River Site (SRS) in 2012-2015. .59
viii
Table 2-13: Concentrations of trace elements (ppm, dry mass) in muscle and liver tissues
of other water birds collected from the Savannah River Site (SRS) in 2012-2015.60
Table 2-14: Correlationsa among trace element concentrations in muscle (above diagonal)
and liver (below diagonal) tissues of all waterfowl/waterbirds collected from the
Savannah River Site (SRS) in 2012-2015 (n=66 muscle, n=70 liver). Only those
elements with more than 50% of the values above detectable limits are included.
Correlations between muscle and liver samples for individual elements are
presented on the diagonal in bold. .........................................................................61
Table 2-15: Descriptive statistics for radiocesium concentrations measured in waterfowl
collected from the Savannah River Site (SRS) in 2012-2015. ...............................62
Table 2-16: Comparisons of whole-body and tissue radiocesium concentrations (Bq/g,
wet mass) in diving ducksa collected from four different locations of the SRS in
2012-2015. .............................................................................................................63
Table 2-17: Monthly allowances of ½ lb. meals for adults and ¼ lb. for children before
exceeding the EPA’s oral reference dose ratings for selenium (Se) and mercury
(Hg) for muscle tissue of wild pigs collected from the Savannah River Site (SRS)
(n=88) and five counties in Georgia (GA) (n=20) 2012-2015. Consumption limits
based on average concentrations are presented with consumption limits based on
the maximum concentration found in an individual in parentheses. Levels of As
were all BDL so consumption limits are not included. ..........................................64
Table 2-18: The monthly allowances of ½ lb. meals for adults and ¼ lb. for children
before exceeding the EPA’s chronic oral reference dose limits for arsenic (As),
selenium (Se), and mercury (Hg) for muscle tissue of diving ducks collected from
ix
the D-Area ash basins (n=24) and other water bodies (n=18) on the Savannah
River Site (SRS) 2012-2015. Consumption limits based on average concentrations
are presented with limits based on the maximum concentration found in an
individual for each trace element in parentheses. Levels of As for Dabbling ducks
were all BDL so consumption limits are not included. ..........................................65
Table 3-1: Tissue and blood concentrations of arsenic (As), selenium (Se), and mercury
(Hg) of each recollected ring-necked duck restricted to the D-Area ash basins
(n=33) on the Savannah River Site (SRS) in the winter of 2014-2015 with
between 3 and 92 days of exposure. ....................................................................106
Table 3-2: Trace elements from ring-necked ducks (n=33) before and after being
restriction to the D-Area ash basins on the Savannah River Site (SRS) for between
3 and 92 days in winter of 2014-2015. Data are shown in approximately 15-day
increments of exposure with the mean±SE concentrations for each element. .....107
Table 3-3: Selenium (Se) linear regression with days of exposure for recollected ring-
necked ducks restricted to the D-Area ash basins (n=33) between 3 and 92 days
on the Savannah River Site (SRS) in the winter of 2014-2015. ..........................109
Table 3-4: Arsenic (As) linear regression with days of exposure for recollected ring-
necked ducks restricted to the D-Area ash basins (n=33) between 3 and 92 days
on the Savannah River Site (SRS) in the winter of 2014-2015. ..........................109
Table 3-5: Mercury (Hg) linear regression with days of exposure for recollected ring-
necked ducks restricted to the D-Area ash basins (n=33) between 3 and 92 days
on the Savannah River Site (SRS) in the winter of 2014-2015. ..........................109
x
Table 3-6: Correlations among trace element concentrations in muscle (above diagonal)
and liver (below diagonal) tissues for recollected ring-necked ducks restricted to
the D-Area ash basins (n=33) between 3 and 92 days on the Savannah River Site
(SRS) in the winter of 2014-2015. Correlations between muscle and liver samples
for individual elements are presented on the diagonal in bold. ...........................110
Table 3-7: The monthly allowances of ½ lb. meals for adults and ¼ lb. for children before
exceeding the EPA’s chronic oral reference dose limits for arsenic (As), selenium
(Se), and mercury (Hg) for muscle tissue of ring-necked ducks collected from the
D-Area ash basins on the Savannah River Site (SRS) after being restricted
between 3 and 92 days of exposure. Consumption limits based on average
concentrations of cooked ducks muscle are presented with limits based on the
maximum concentration found in an individual for each trace element in
parentheses. ..........................................................................................................111
Table 4-1: Ecological half-life estimates for species from Pond B or Par Ponda on the
Savannah River Site (SRS). For the current study American coots were restricted
to Pond B for between 33 and 173 days of exposure to radiocesium in that system.147
Table 4-2: Descriptive statistics for a random sampling of American coots and ring-
necked ducks that were trapped from L-Lake and whole-body counted for
radiocesium prior to release onto Pond B on the Savannah River Site (SRS) over
the winter of 2013-2015. ......................................................................................148
Table 4-3: Descriptive statistics for radiocesium concentrations of American coots and
ring-necked ducks that were released to Pond B on the Savannah River Site (SRS)
for between 33 and 173 days of exposure before being collected. ......................149
xi
LIST OF FIGURES
Page
Figure 2-1: Savannah River Site (SRS) locations targeted for sample collections (wild
pigs, squirrels, waterfowl/waterbirds) for trace elements and radiocesium
quantification in 2012-2015, included the D-Area ash basins, Fourmile Branch,
Tim’s Branch, Pond A/R-Canal, Pond B, and L-Lake. .........................................66
Figure 3-1: D-Area ash basins on the Savannah River Site (SRS), SC. Basin 1, the largest
basin is partially filled in and has extensively revegetated. The smaller enclosed
wetland formed by revegetation in this basin was utilized as the release and
exposure area for the ring-necked ducks in this study in winter of 2014-2015. ..112
Figure 3-2: Muscle concentrations of arsenic (As), selenium (Se), mercury (Hg), and days
of exposure of recollected ring-necked duck restricted to the D-Area ash basins
(n=33) on the Savannah River Site (SRS) in the winter of 2014-2015 between 3
and 92 days of exposure. ......................................................................................113
Figure 3-3: Liver concentrations of arsenic (As), selenium (Se), mercury (Hg), and days
of exposure of recollected ring-necked duck restricted to the D-Area ash basins
(n=33) on the Savannah River Site (SRS) in the winter of 2014-2015 between 3
and 92 days of exposure. ......................................................................................114
Figure 3-4: Blood concentrations of arsenic (As), selenium (Se), mercury (Hg), and days
of exposure of recollected ring-necked duck restricted to the D-Area ash basins
xii
(n=33) on the Savannah River Site (SRS) in the winter of 2014-2015 between 3
and 92 days of exposure. ......................................................................................115
Figure 4-1: The Par Pond Reservoir system on the Savannah River Site (SRS) that
includes P- and R-reactors with depictions of canals that carried the radionuclide
contaminated cooling water to Ponds B and C and Par Pond during several reactor
releases. ................................................................................................................150
Figure 4-2: Historical and current data for whole-body radiocesium concentrations in
American coots collected from Pond B on the Savannah River Site between the
winters of 1975-1976 and 2014-2015; a.) includes a dashed line for the whole-
body equivalent (0.324 Bq/g) to the European Economic Community limit for
radiocesium in fresh meat (0.600 Bq/g) and b) is a linear regression of natural log-
transformed data from collections between winters 1986-1987 and 2014-2015,
estimates utilized in ecological half-life calculations. .........................................151
Figure 4-3: Whole-body radiocesium concentrations in a.) American coots and b.) ring-
necked ducks collected from Pond B on the Savannah River Site (SRS) after
exposure between 32 and 173 days. Day 0 whole-body concentrations are counts
done on live-captured birds from L-Lake before release onto Pond B. Solid lines
show non-linear fits to the data and dashed lines represent our calculated whole-
body equivalent (0.324 Bq/g) to the European Economic Community limit for
radiocesium in fresh meat (0.600 Bq/g). ..............................................................152
Figure 4-4: Radiocesium concentrations in muscle and liver tissue graphed against
whole-body concentrations of a.) American coots and b.) ring-necked ducks
xiii
collected from Pond B on the Savannah River Site (SRS) after exposure between
32 and 173 days. Lines are linear fits with intercepts constrained to zero. .........153
1
CHAPTER 1
INTRODUCTION AND LITERATURE REVIEW
Wildlife face numerous anthropogenic pressures globally, especially direct and
indirect threats from human development and the resulting reduction and degradation of
remaining native habitats. In particular, anthropogenic activities have released numerous
pollutants into the environment, resulting in potential adverse health effects in both
wildlife and human consumers of wildlife. Wildlife can be exposed to contaminants via
physical contact with, or ingestion of, pollutants, raising a concern because many wildlife
species are harvested for food and these pollutants could be passed on to human
consumers. The continued exposure through ingestion of contaminated food can result in
the accumulation of contaminants in the tissues of exposed individuals, resulting in
greater body burdens in higher tropic level consumers through biomagnification (USEPA
2012). Because humans are apex predators in many systems, it is important to elucidate
patterns of bioaccumulation of contaminants within food webs.
Hunting and consumption of wildlife is important culturally but also as source of
food globally, including in the U.S. In the last U.S. Fish and Wildlife census in 2011
there were 13.7 million hunters, an increase of 9 percent from 2006 (USFWS and USCB
2011). Despite the potential for many pollutants to bioaccumulate or biomagnify within
food webs, game animals are not subject to the same scrutiny and testing as farm raised
livestock and poultry, leaving a gap in knowledge of contaminant exposure resulting
2
from consumption of game, like recent concerns showing increased levels of lead in
hunters using lead ammunition to hunt game animals (Pain et al. 2010).
Human generated pollution can affect wildlife worldwide but game species with
large home ranges or migratory behaviors are of particular interest for monitoring
(Taggart et al. 2011). Of specific concern are waterfowl and waterbirds, which can travel
hundreds to thousands of miles during migration (Kennamer 2003, Cristol et al. 2012,
Conder and Arblaster 2016). Despite the fact that the U.S. has lost approximately 52% of
its original wetlands over the last few centuries, populations of many waterfowl species
have been increasing in recent decades and currently are 43% higher than the long-term
average since population estimates first originated in 1955 (Dahl 2000, USFWS 2014).
As conversion of wetlands continues, increasing numbers of birds will need to use bodies
of water that have a history of anthropogenic disturbances (Foley and Batcheller 1988).
In particular, use of aquatic habitats contaminated with toxic substances such as trace
elements, PCBs, organochlorines, and radiocesium (137Cs) is a serious concern. Wildlife
using contaminated sites may suffer deleterious effects and potential exists for human
exposure to contaminants through consumption of contaminated meat even far from sites
of contamination (Brisbin et al. 1973; Fendley et al. 1977; Brisbin and Vargo 1982; Foley
and Batcheller 1988; Kennamer et al., 1998; Sajwan et al., 2009; Cristol et al. 2012,
Kalisinska et al., 2013). However, despite increases in waterfowl numbers and the
potential for human exposure, there is limited data on the extent and types of
contamination waterfowl could be passing on to hunters, as well as how long waterfowl
have to forage in a contaminated area before they become a risk to hunters.
3
There are many ways in which wildlife can be exposed to contaminants in the
environment. Of increasing interest is exposure to trace elements found in coal
combustion wastes (CCW). Coal fly ash effluent discharged into Belews Lake and Hyco
Reservoir in North Carolina and in Martin Creek Reservoir in Texas has resulted in trace
element accumulation in aquatic biota (Lemly 1996, 2002). Further, studies associated
with a large fly ash effluent spill from the Kingston Plant in Tennessee showed that even
terrestrial animals can be exposed to elevated levels of elements found in fly ash
deposited in settling basins (Ruhl et al. 2009). Coal-fly ash contains many trace elements,
such as Al, As, Cd, Cr, Cu, Fe, Hg, Mn, Ni, Se, and Zn, which can leach out of the ash
into the water and combine with sediments (Cherry and Guthrie 1977; Evans and Giesy
1978; Alberts et al. 1985; Sandhu et al. 1993). Trace elements such as As, Se, and Hg are
of special concern in ash basins because of the potentially deleterious effects to human
consumers of exposed wildlife (Luther 2010).
One of the most toxic metals affecting organisms is Hg (Diez et al. 2008), which
is often found within coal fly ash. Hg accumulates in tissues and organs such as muscle,
kidney, liver, and the brain (Graeme and Pollack 1998). Methylmercury, (MeHg) the
most toxic form of Hg, can cause severe neurological damage to humans and wildlife
(Grandjean et al. 1999, Clarkson et al. 2003). Acute effects of Hg at high concentrations
have been well-documented, but there are also adverse effects identified at lower tissue
concentrations, representing chronic exposure (Wolfe et al. 1998). In humans Hg affects
neurobehavioral functioning (Goyer 1991); high doses cause problems with coordination
of movement, muscle weakness, as well as sensibility, language, vision and hearing
(Duchesne et al. 2004). Exposure to Hg is especially important for sensitive populations
4
like expectant mothers and children. Tracking the transfer to humans through
consumption of fish and game is vital to prevent overexposure to this pollutant (USEPA
2014).
Selenium is often found in CCW settling basins where it is primarily found in
sediments (Lemy 2002, Luther 2010) but is readily transferred through food webs.
Therefore, wildlife can be exposed to Se through consumption of vegetation grown from
contaminated sediments, or incidental ingestion of sediments while foraging, as well as
through consumption of other species that are bioaccumulating contaminants (Heinz
1996). In humans and wildlife exposure to Se may result in disruption of endocrine
function, neurotoxicity, hepatotoxicity, and effects on reproductive capabilities (USEPA
1999). Similarly, As is often found in fly ash (Ruhl et al. 2009, Luther 2010) and also
poses a threat to wildlife and human health, resulting in increasing blood pressure leading
to heart attacks. As exposure can also result in increased risk of skin, lung, bladder, and
kidney cancer in humans (USEPA 1998).
Interactive effects among trace elements can influence the rates of uptake,
elimination, and sometimes the area of deposition within the body (Lorentzen et al. 1998,
Kehrig et al. 2009, García-Barrera et al. 2012). In mallards (Anas platyrhynchos) it
appears Hg and As act antagonistically with Se (Hoffman et al. 1992, Heinz and Hoffman
1998). In mallard embryos MeHg and selenomethionine were cumulative or synergistic
in toxicity in regards to survival and teratogenesis (Hoffman and Heinz 1998). Moreover,
the adverse effects from Se exposure such as mortality, impaired growth,
histopathological lesions, and resultant oxidative stress have been found to be
5
ameliorated by interactions with Hg and As in laboratory experiments (Heinz 1979,
Heinz 1996, Heinz and Hoffman 1998).
Radioactive materials from nuclear activities such as atmospheric fallout from
nuclear weapons testing and nuclear accidents like those in Chernobyl and Fukushima are
another source of contamination that can affect wildlife (UNSCEAR 1988; Gudiksen et
al. 1989; Mishra 1990; Bennet 1995; Dreicer et al. 1996; Masson et al. 2011; Evangeliou
et al. 2013; Masson et al. 2013, Tagami and Uchida 2013).Within the U.S. there are
dozens of Department of Energy (DOE) sites where nuclear weapons activities have
contaminated the environment with radionuclides (e.g. Willard 1960, Fitzner and Rickard
1975, Halford et al. 1981). For example, radionuclides and chromium (Cr) discharged
into the Columbia River in Washington from the Hanford Nuclear Reservation
contaminated breeding grounds of Chinook salmon (Farag 2006).
Release of radionuclides can have significant effects on the environment and
humans for many years given the long half-lives of many isotopes (Skuterud et al. 2005,
Hinton et al. 2007, Christodouleas et al. 2011). In particular, 137Cs has a radioactive half-
life of 30.2 years and is slowly mobilized from sediments (Evans et al. 1983).
Radiocesium is of biological importance because its chemical behavior is similar to
potassium and thus can accumulate in skeletal muscle, particularly in potassium deficient
environments (Brisbin 1991).
Clearly, there is a growing importance to elucidate the potential health effects of
contaminant exposure to wildlife and the levels of contaminant accumulation within
consumable tissues of game species to assess human risks. Despite potential risks to
wildlife and human consumers, there are limited data available on contaminant levels
6
within many common game species in North America (Wolkers et al. 1994; Kennamer et
al. 1998; Braune and Malone 2006; Dvořák et al. 2010; Taggart et al. 2011, Tagami and
Uchida 2013). Thus, these data could have applications in regulating the allowable
release and remediation efforts of pollutants, as well as the consumption of game in
specific areas. Wildlife of higher trophic levels or that spend a greater portion of the year
in contaminated areas would be expected to have higher levels of trace elements or
radiocesium and thus be a greater potential risk to human consumers, but knowledge of
how residence times influence bioaccumulation in animals also is lacking (Hall et al.
2009 and Vest et al. 2009).
The goals of this research were to elucidate levels of common trace elements in
CCW and radiocesium in several common game species, the accumulation rates of these
contaminants, and how observed contaminant burdens relate to human consumption risks.
To accomplish this, several popular game species in the U.S. [wild pigs (Sus scrofa), gray
squirrels (Sciurus carolinensis), and waterfowl/waterbirds] were collected between 2012
and 2015 from the Savannah River Site (SRS), a DOE property in South Carolina, which
contains uncontaminated habitats and areas of known contamination, allowing for
collection of animals across a diversity of contaminated and uncontaminated areas.
Various tissue samples were taken from the game animals and tested for trace elements
and/or radiocesium.
Specifically, in Chapter 2 I quantified trace element and radiocesium levels in
free-ranging game animals inhabiting both contaminated and reference areas and
compared observed contaminant concentrations of free-ranging game animals to
recommended limits for human consumption and levels known to cause deleterious
7
effects to the game animals. In Chapter 3 my objectives were to 1) quantify trace element
uptake in blood, muscle, and liver tissues over known periods of time by waterfowl
exposed in situ to a coal ash settling basin and investigate potential accumulation rates, 2)
develop a model to predict muscle/liver burdens based on concentrations in blood as a
potential non-destructive sampling method and test the performance of the model against
a subset of our data, and 3) calculate human consumption limits based on concentrations
of recognized elements of human health concern (As, Se, and Hg) over known time
periods of exposure. For Chapter 4 I quantified uptake of radiocesium at multiple time
points by migrating American coots (Fulica americana) and ring-necked ducks (Aythya
collaris) subsequent to translocation to a radiocesium contaminated reservoir and
compared observed accumulation patterns between the species. I also calculated the
ecological half-life for radiocesium in American coots at this site by comparing whole-
body burdens to historical data collected from the same location. Collectively, this
research seeks to present levels of contaminants for several common mammalian and
avian game species and how certain waterfowl/waterbird species accumulate
contaminants over time when present in polluted ecosystems. These results will have
implications for wildlife inhabiting sites with similar contaminants around the world.
LITERATURE CITED
Alberts, J. J., and T. J. Dickson. 1985. Organic carbon and cation associations in humic
material from pond water and sediment. Organic Geochemistry 8:55-64.
8
Bennett, B. G. 1995. Exposures from worldwide releases of radionuclides.
Environmental Impact of Radioactive Releases. International Atomic Energy
Agency (IAEA), Vienna Austria, pp. 3-12.
Braune, B. M., and B. J. Malone. 2006. Organochlorines and trace elements in upland
game birds harvested in Canada. Science of the Total Environment 363:60-69.
Brisbin Jr., I. L., R. A. Geiger, and M. H. Smith. 1973. Accumulation and redistribution
of radiocesium by migratory waterfowl inhabiting a reactor cooling reservoir.
Environmental behavior of radionuclides released in the nuclear industry,
International Atomic Energy Agency symposium (IAEA-SM-172/72) Vienna,
Austria, pp. 373-384.
Brisbin Jr., I. L., and M. J. Vargo. 1982. Four-year declines in radiocesium
concentrations of American coots inhabiting a nuclear reactor cooling reservoir.
Health Physics 43:266-269.
Brisbin Jr., I. L. 1991. Avian radioecology. Current Ornithology, Volume 8. (Ed. D.M.
Power) Plenum Publishing Corporation, New York, NY, pp. 69-140.
Cherry, D. S., and R. K. Guthrie. 1977. Toxic metals in surface waters from coal ash.
Water Resources Bulletin 13:1227-1236.
Christodouleas, J. P., R. D. Forrest, C. G. Ainsley, Z. Tochner, S. M. Hahn, and E.
Glatstein. 2011. Short-term and long-term health risks of nuclear-power-plant
accidents. New England Journal of Medicine 364:2334-2341.
Clarkson, T. W., L. Magos, and G. J. Meyers. 2003. Human exposure to mercury: The
three modern dilemmas. Journal of Trace Elements in Experimental Medicine
16:321-343.
9
Conder, J. M., and J. A. Arblaster. 2016. Development and use of wild game
consumption rates in human health risk assessments. Human and Ecological Risk
Assessment 22:251-264.
Cristol, D. A., L. Savoy, D. C. Evers, C. Perkins, R. Taylor, and C. W. Varian-Ramos.
2012. Mercury in waterfowl from a contaminated river in Virginia. Journal of
Wildlife Management 76:1617-1624.
Dahl, T. E. 2000. Report to Congress on the status and trends of wetlands in the
conterminous United States 1986 to 1997. United States Department of Interior,
Fish and Wildlife Service, Washington, DC, pp. 82.
Diez, S., C. Barata, and D. Raldua. 2008. Exposure to mercury: A critical assessment of
adverse ecological and human health effects. Trace elements as contaminants and
nutrients: Consequences in ecosystems. (Ed. M.N.V. Prasad) John Wiley & Sons,
Inc., Hoboken, NJ.
Dreicer, M., A. Aarkrog, R. Alexakhin, L. Anspaugh, N. P. Arkhipov, and K. J.
Johansson. 1996. Consequences of the Chernobyl accident for the natural and
human environments. One Decade after Chernobyl: Summing up the
Consequences of the Accident. International Atomic Energy Agency (IAEA),
Vienna, Austria, pp. 319-361.
Duchesne, J. F., B. Levesque, D. Gauvin, B. Braune, S. Gingras, and E. Dewailly. 2004.
Estimating the mercury exposure dose in a population of migratory bird hunters in
the St. Lawrence River region, Quebec, Canada. Environmental Research 95:207-
217.
10
Dvořák, P., P. Snášel, and K. Beňová. 2010. Transfer of radiocesium into wild boar meat.
Acta Veterinaria Brno 79:85-91.
Evangeliou, N., Y. Balkanski, A. Cozic, and A. P. Møller. 2013. Global Transport and
deposition of 137Cs following the Fukushima nuclear power plant accident in
Japan: Emphasis on Europe and Asia using high–resolution model versions and
radiological impact assessment of the human population and the environment
using interactive tools. Environmental Science and Technology 47:5803-5812.
Evans, D.W., and J.P. Giesy Jr. 1978. Trace metal concentrations in a stream-swamp
system receiving coal ash effluent. Ecology and Coal Resource Development Vol.
2 (Ed. M.K. Wali). Proceedings of the International Congress on Energy and the
Ecosystem, Grand Fork, ND, pp. 782-790.
Evans, D. W., J. J. Alberts, and R. A. Clark. 1983. Reversible ion-exchange fixation of
cesium-137 leading to mobilization from reservoir sediments. Geochimica et
Cosmochimica Acta 47:1041–1049.
Farag, A. M., D. D. Harper, L. Cleveland, W. G. Brumbaugh, and E. E. Little. 2006. The
potential for chromium to affect the fertilization process of chinook salmon
(Oncorhynchus tshawytscha) in the Hanford Reach of the Columbia River,
Washington, USA. Archives of Environmental Contamination and Toxicology
50:575-579.
Fendley, T. T., M. N. Manlove, and I. L. Brisbin Jr. 1977. The accumulation and
elimination of radiocesium by naturally contaminated wood ducks. Health Physics
32: 415-422.
11
Fitzner, R. E., and W. H. Rickard. 1975. Avifauna of waste ponds ERDA Hanford
Reservation, Benton County, Washington. Battelle Pacific Northwest Labs
Report.
Foley, R. E., and G. R. Batcheller. 1988. Organochlorine contaminants in common
goldeneye wintering on the Niagara River. Journal of Wildlife Management
52:441-445.
García-Barrera, T., J. L. Gómez-Ariza, M. González-Fernández, F. Moreno, M. A.
García-Sevillano, and V. Gómez-Jacinto. 2012. Biological responses related to
agonistic, antagonistic and synergistic interactions of chemical species. Analytical
and Bioanalytical Chemistry 403:2237-2253.
Goyer, R. A. 1991. Toxic effects of metals. Casarett and Doull’s Toxicology, Fourth
edition. (Eds. C.D. Klassen, M.O. Amdur, J. Doull) Macmillan, New York, NY,
pp. 623-680.
Graeme, K. A., and C. V. Pollack. 1998. Heavy metal toxicity, part I: Arsenic and
mercury. Journal of Emergency Medicine 16:45-56.
Grandjean, P., R. F. White, A. Nielsen, D. Cleary, and E. C. Santos. 1999.
Methylmercury neurotoxicity in Amazonian children downstream from gold
mining. Environmental Health Perspectives 107:587-591.
Gudiksen, P. H., T. F. Harvey, and R. Lange. 1989. Chernobyl source term, atmospheric
dispersion, and dose estimation. Health Physics 57:697-706.
Hall, B. D., L. A. Baron, and C. M. Somers. 2009. Mercury concentrations in surface
water and harvested waterfowl from the prairie pothole region of Saskatchewan.
Environmental Science and Technology 43:8759-8766.
12
Halford, D. K., J. B. Millard, and D. O. Markham. 1981. Radionuclide concentrations in
waterfowl using a liquid radioactive waste disposal area and the potential
radiation dose to man. Health Physics 40:173-181.
Heinz, G. H. 1979. Methylmercury: Reproductive and behavioral effects on three
generations of mallard ducks. Journal of Wildlife Management 43:394-401.
Heinz, G. H. 1996. Selenium in birds. Environmental Contaminants in Wildlife:
Interpreting Tissue Concentrations. (Eds. W. N. Beyer, G.H. Heinz, and A. W.
Redmon-Norwood) CRC Press, Boca Raton, FL, pp. 447-458.
Heinz, G. H., and D. J. Hoffman. 1998. Methylmercury chloride and selenomethionine
interactions on health and reproduction in mallards. Environmental Toxicology
and Chemistry 17:139-145.
Hinton, T. G., R. Alexakhin, M. Balonov, N. Gentner, J. Hendry, B. Prister, P. Strand,
and D. Woodhead. 2007. Radiation-induced effects on plants and animals:
Findings of the United Nations Chernobyl Forum. Health Physics 93:427-440.
Hoffman, D. J., C. J. Sanderson, L. J. LeCaptain, E. Cromartie, and G. S. Pendleton.
1992. Interactive effects of arsenic, selenium, and dietary protein on survival,
growth, and physiology in mallard ducklings. Archives of Environmental
Contamination and Toxicology 20:288-294.
Hoffman, D. J. and G. H. Heinz. 1998. Effects of mercury and selenium on glutathione
metabolism and oxidative stress in mallard ducks. Environmental Toxicology and
Chemistry 17:161-166.
Kalisinska, E., D. I. Kosik-Bogacka, P. Lisowski, N. Lanocha, and A. Jackowski. 2013.
Mercury in the body of the most commonly occurring European game duck, the
13
Mallard (Anas platyrhynchos L. 1758), from Northwestern Poland. Archives of
Environmental Contamination and Toxicology, 64:583-593.
Kehrig, H. D. A., T. G. Seixas, E. A. Palermo, A. P. Baêta, C. W. Castelo-Branco, O.
Malm et al. 2009. The relationships between mercury and selenium in plankton
and fish from a tropical food web. Environmental Science and Pollution Research
16:10-24.
Kennamer, R. A., I. L. Brisbin Jr., C. D. McCreedy, and J. Burger. 1998. Radiocesium in
Mourning Doves: Effects of a contaminated reservoir drawdown and risk to
human consumers. Journal of Wildlife Management 62:497-508.
Kennamer, R. A. 2003. Recoveries of ring-necked ducks banded on the U.S. Department
of Energy’s Savannah River Site, South Carolina. The Oriole 68:8-14.
Lemly, A. D. 1996. Selenium in aquatic organisms. Environmental Contaminants in
Wildlife: Interpreting Tissue Concentrations. (Eds. W. N. Beyer, G. H. Heinz, and
A. W. Redmon-Norwood) CRC Press, Boca Raton, FL, pp. 427-455.
Lemly, A. D. 2002. Symptoms and implications of selenium toxicity in fish: the Belews
Lake case example. Aquatic Toxicity 57:39-49.
Lorentzen, M., A. Maage, and K. Julshamn. 1998. Supplementing copper to a fish meal
based diet fed to Atlantic salmon parr affects liver copper and selenium
concentrations. Aquaculture Nutrition 4:67-72.
Luther, L. 2010. Regulating Coal Combustion Waste Disposal: Issues for Congress.
Congressional Research Service. CRS report for Congress, 7-5700, R41341, pp.
1-22. http://www.fas.org/sgp/crs/misc/R41341.pdf. Accessed 29 January 2016.
14
Masson, O., A. Baeza, J. Bieringer, K. Brudecki, S. Bucci, M. Cappai, F. P. Carvalho et
al. 2011. Tracking of airborne radionuclides from the damaged Fukushima Dai-
ichi nuclear reactors by European networks. Environmental Science and
Technology 45:7670-7677.
Masson, O., W. Ringer, H. Malá, P. Rulik, M. Dlugosz-Lisiecka, K. Eleftheriadis, O.
Meisenberg, A. De Vismes-Ott, and F. Gensdarmes. 2013. Size distributions of
airborne radionuclides from the Fukushima nuclear accident at several places in
Europe. Environmental Science and Technology 47:10995-11003.
Mishra, U. 1990. Comparison of radionuclides levels from the Chernobyl reactor accident
and from global fallout. Journal of Radioanalytical and Nuclear Chemistry
138:119-125.
Pain, D.J., R. L. Cromie, J. Newth, M. J. Brown, E. Crutcher, P. Hardman, L. Hurst, R.
Mateo, A. A. Meharg, A. C. Moran, and A. Raab. 2010. Potential hazard to
human health from exposure to fragments of lead bullets and shot in the tissues of
game animals. PLoS One 5:10315.
Ruhl, L., A. Vengosh, G. S. Dwyer, H. Hsu-Kim, A. Deonarine, M. Bergin, and J.
Kravchenko. 2009. Survey of environmental and health impacts in the immediate
aftermath of the coal ash spill in Kingston, Tennessee. Environmental Science and
Technology 43:6326-6333.
Sandhu, S. S., G. L. Mills, and K. S. Sajwan. 1993. Leachability of Ni, Cd, Cr, and As
from coal ash impoundments of different ages on the Savannah River Site. Trace
Elements in Coal and Coal Combustion Residues (Eds. R. F. Keefer and K. S.
Sajwan). Lewis Publishers, Boca Raton, FL.
15
Sajwan, K. S., K. S. Kumar, S. Kelley, and B. G. Loganathan. 2009. Deposition of
organochlorine pesticides, PCBs (Aroclor 1268), and PBDEs in selected plant
species from a superfund site at Brunswick, Georgia, USA. Bulletin of
Environmental Contamination and Toxicology 82:444-449.
Skuterud, L., E. Gaare, I. M. Eikelmann, K. Hove, and E. Steinnes. 2005. Chernobyl
radioactivity persists in reindeer. Journal of Environmental Radioactivity 83:231-
252.
Tagami, K., and S. Uchida. 2013. Radiocesium concentration change in game animals:
Use of food monitoring data. Waste Management Conference: International
collaboration and continuous improvement INIS-US--13-WM-13168.
Taggart, M. A., M. M. Reglero, P. R. Camarero, R. Mateo. 2011. Should legislation
regarding maximum Pb and Cd levels in human food also cover large game meat?
Environmental International 37:18-25.
United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR).
1988. Sources, effects, and risks of ionizing radiation. United Nations, NY, pp.
647.
United States Environmental Protection Agency (USEPA). 1998. Integrated risk
information system (IRIS) on arsenic. National Center for Environmental
Assessment, Office of Research and Development, Washington, DC.
http://www.epa.gov/ttnatw01/hlthef/arsenic.html#ref5. Accessed 30 January
2015.
United States Environmental Protection Agency (USEPA). 1999. Integrated risk
information system (IRIS) on selenium and compounds. National Center for
16
Environmental Assessment, Office of Research and Development, Washington,
DC. http://www.epa.gov/airtoxics/hlthef/selenium.html#ref4. Accessed 30
January 2015.
United States Environmental Protection Agency (USEPA). 2012. Waste and cleanup risk
assessment glossary. U.S. Environmental Protection Agency Office of Solid
Waste Emergency Response, Washington, DC.
http://www.epa.gov/oswer/riskassessment/glossary.html. Accessed 16 October
2014.
United States Environmental Protection Agency (USEPA). 2014. Mercury health effects.
Washington, DC. http://www.epa.gov/mercury/effects.htm. Accessed 3
November 2014.
United States Fish and Wildlife Service (USFWS) and United States Census Bureau
(USCB). 2011. 2011 National survey of fishing, hunting and wildlife associated
recreation. U.S. Department of the Interior, Washington, DC.
https://www.census.gov/prod/2012pubs/fhw11-nat.pdf. Accessed 09 February
2015.
United States Fish and Wildlife Service (USFWS). 2014. Waterfowl population status,
2014. U.S. Department of the Interior, Washington, DC.
http://www.fws.gov/migratorybirds/NewReportsPublications/PopulationStatus/W
aterfowl/StatusReport2014.pdf. Accessed 14 October 2014.
Vest, J. L., M. R. Conover, C. Perschon, J. Lift, and J. O. Hall. 2009. Trace element
concentrations in wintering waterfowl from the Great Salt Lake, Utah. Archives
of Environmental Contamination and Toxicology 56:302-316.
17
Willard, W. K. 1960. Avian uptake of fission products from an area contaminated by
low-level atomic wastes. Science 132:148-150.
Wolfe, M. F., S. Schwarzbach, and R. A. Sulaiman. 1998. Effects of Mercury on wildlife:
A comprehensive review. Environmental Toxicology and Chemistry 17:146-160.
Wolkers, H., T. Wensing, and G. W. Bruinderink. 1994. Heavy metal contamination in
organs of red deer (Cervus elaphus) and wild boar (Sus scrofa) and the effect on
some trace elements. Science of the Total Environment 144:191-199.
18
CHAPTER 2
TRACE ELEMENTS AND RADIOCESIUM IN GAME SPECIES AND HUMAN
CONSUMPTION RISKS 1
____________________
1 Oldenkamp, R. E., A. L. Bryan Jr., R. A. Kennamer, S. C. Webster, and J. C. Beasley. To be submitted to the Journal of Wildlife Management.
19
ABSTRACT
Mercury (Hg), Selenium (Se), and Arsenic (As) found in coal combustion wastes
(CCW) and radionuclides released from anthropogenic activities are an environmental
and human health concern. Despite the widespread harvest and consumption of wildlife
by recreational hunters, game species are not subject to the same safety testing as
livestock and thus there are few data available regarding contaminant concentrations in
many commonly harvested wildlife. We sampled wild pigs (Sus scrofa), gray squirrels
(Sciurus carolinensis), and waterfowl from uncontaminated habitats as well as areas of
known contamination and quantified levels of trace elements and radiocesium in muscle
and liver tissues for assessment of potential human health risks from the consumption of
game. Our results revealed substantive variability in contamination burdens within and
among species, tissue types, elements, and sampling locations, likely reflecting
differences in resource selection, diet, and behavior. Species collected at a CCW ash
basin consistently had elevated levels of trace elements, particularly Se, suggesting CCW
may be an important pathway for wildlife and subsequently human exposure to this
element. Similarly, we observed elevated concentrations of radiocesium in individuals
from locations with histories of operational releases of radionuclides. The majority of
tissue samples analyzed were below levels known to adversely affect wildlife health and
radiocesium levels were below established EEC limits for human consumption.
Waterfowl consistently had elevated levels of several measured elements of interest (Se
and Hg), especially individuals collected from areas with known contamination. Given
the high levels of trace element burdens we observed in waterfowl collected from the ash
basin site, and the common occurrence of similar surface impoundments in the U.S.
20
additional studies are needed to more clearly elucidate potential risks to both wildlife and
waterfowl hunters.
INTRODUCTION
Anthropogenic pollution can come from a variety of sources including
atmospheric deposition of metals or releases of wastes from factories, run-off of
pesticides and herbicides from farms, and burning of fossil fuels. The U.S. Environmental
Protection Agency (EPA) cites coal and oil-fired electric generating units as the dominant
emitters of mercury, acid gases, and many toxic metals in the U.S. (USEPA 2012a). As
of 2015, in the U.S. alone the EPA identified more than 500 coal power facilities using
735 surface impoundments to store coal combustion waste (CCW; USEPA 2015).
Furthermore, coal use is on the rise in developing countries, making CCW important at a
global scale (Humphries, 1999). The millions of tons of CCW deposited into surface
impoundments across the U.S. (Luther 2010a) contain trace elements that the EPA
considers environmental and human health risks and represent potential pathways for
contaminant exposure for wildlife and human consumers of wildlife (Rowe et al. 2002,
Luther 2010a).
Surface impoundments (hereafter referred to as ash basins) used to store CCW
represent potential habitat for foraging and reproduction for many wildlife species,
especially birds and amphibians, which can result in trace element exposure and
bioaccumulation (Yudovich and Ketris 2005a,b; Yudovich and Ketris 2006; Reash 2012;
Otter et al. 2012). Wildlife can be affected through the consumption of contaminated
water, sediment/soil, and/or food (USEPA 2012b) and examples of trace element
bioaccumulation in wildlife exposed to CCW are well documented (Dorman et al. 2010,
21
Lemly and Skorupa 2012, Ruhl et al. 2012, Mayfield et al. 2013, Rice et al. 2014). For
example, coal fly ash effluent discharged into Belews Lake and Hyco Reservoir in North
Carolina, in Martin Creek Reservoir in Texas, and the largest ash release in U.S. history
into the Clinch and Emory Rivers in Tennessee have resulted in trace element
accumulation in aquatic and some terrestrial biota (Lemly 1996, 2002; Ruhl et al. 2009;
Beck et al. 2013; Van Dyke et al. 2013; Beck at al. 2015).
Although numerous trace elements exist within fly ash that accumulate in tissues
of exposed wildlife, Mercury (Hg), Selenium (Se), and Arsenic (As) are of particular
concern for environmental and human health (Luther 2010b). Exposure to toxicants
found in coal ash such as As, cadmium (Cd), lead (Pb), and Hg can impact an animals’
central nervous system, cardiovascular, hematopoietic, gastrointestinal, urinary, and
reproductive systems (Chmielnicka et al. 1989, Domingo 1994, Hughes et al. 2011). In
particular, effects of Hg at high concentrations have been well-documented, but even
chronic low-dose exposure can have severe adverse effects (Wolfe et al. 1997). In
humans, Hg affects neurobehavioral functioning (Goyer 1991); high doses cause
problems with coordination of movement, muscle weakness, sensibility, language, vision
and hearing (Duchesne et al. 2004), and is of particular concern for expectant mothers
and children (USEPA 2014). Exposure to essential elements such as Cr, copper (Cu), Se,
and zinc (Zn) in high doses can affect hematopoietic, gastrointestinal, and reproductive
systems of animals (Ohlendorf and Flemming 1988, Domingo 1994, Eisler 1998) and in
humans exposure to excessive Se can disrupt endocrine function as well as cause
neurotoxicity and hepatotoxicity (USEPA 1999b). Similarly, accumulation of As can
22
increase blood pressure, leading to heart attacks, and can result in increased risk of skin,
lung, bladder, and kidney cancer (USEPA 1998).
In addition to contaminant exposure through CCW, wildlife also can be exposed
to radioactive materials from nuclear activities such as atmospheric fallout from nuclear
weapons testing and nuclear accidents (e.g., Chernobyl and Fukushima) that released
large amounts of radionuclides (e.g., 131I, 137Cs, 134Cs) into the local environment, as well
as dispersed and deposited radioactive material across the Northern Hemisphere
(UNSCEAR 1988, Gudiksen et al. 1989, Mishra 1990, Bennet 1995, Dreicer et al. 1996,
Masson et al. 2011, Evangeliou et al. 2013, Masson et al. 2013). Within the U.S. there
also are dozens of sites that have been impacted by the U.S. Department of Energy’s
(DOE) nuclear weapons activities. For example, radionuclides and chromium (Cr)
discharged into the Columbia River in Washington, U.S.A. contaminated breeding
grounds of Chinook salmon (Farag 2006). Although release of radionuclides into the
environment has happened at a lower frequency than CCW releases, these events can
have a profound effect on the environment and humans for many years given the long
physical half-lives of many radionuclides (Skuterud et al. 2005, Hinton et al. 2007,
Christodouleas et al. 2011).
As of 2011 there were 13.7 million hunters within the U.S. (USFWS and USCB
2011), the vast majority of whom pursue game for consumption. Consumption of wild
game meat has a reputation for being healthier than many farmed raised meats, but wild
game can present a potential route of contaminant exposure to human consumers as game
animals are not subject to the same testing as livestock (Duchesne et al. 2004, Cristol et
al. 2012, Conder and Arblaster 2016). Wildlife can be exposed to contaminants directly at
23
point sources (e.g., through foraging or other behaviors resulting in incidental sediment
ingestion; Neely 1980, Boening 2000). Moreover, game animals with large home range
sizes (e.g., wild pigs – Sus scrofa, white-tailed deer – Odocoileus virginianus) or
migratory behavior (e.g., ducks) can transport contaminants extensive distances from
point sources, exposing unknowing hunters, their families, and recipients of donated
game meat to contaminants accumulated outside the vicinity of their hunting area
(Kennamer 2003, Taggart et al. 2011, Cristol et al. 2012, Conder and Arblaster 2016).
Despite the potential exposure of human hunters to contaminants through
consumption of game, there are limited data available on contaminant levels within many
common game species in North America (Wolkers et al. 1994, Kennamer et al. 1998,
Braune and Malone 2006, Dvořák et al. 2010, Taggart et al. 2011). Moreover, although
laboratory studies are typically performed to identify lethal doses of contaminants and
measure morbidity, the complex exposure pathways and consequent levels of
contaminants found in natural systems require in situ studies to elucidate accumulation
patterns of trace elements and radiocesium. Such data are essential to inform public
health officials and wildlife managers about actual risks to the public and/or the animals
utilizing those systems.
For this study, we collected several common, popular game species in the U.S.
[wild pigs, gray squirrels (Sciurus carolinensis), and waterfowl] from the Savannah River
Site (SRS), a DOE property in South Carolina with both uncontaminated habitats and
areas of known contamination, to quantify trace element and radiocesium burdens in
consumable tissues. Our objectives were to 1) quantify trace element and radiocesium
levels in game animals inhabiting both contaminated and reference areas and to 2)
24
compare observed contaminant burdens to recommended limits for human consumption
and levels known to cause deleterious effects to the game animals themselves. These
results have implications for wildlife inhabiting sites with similar contaminants across the
globe.
METHODS
Study Area
This research primarily occurred on the Savannah River Site (SRS), a ~800 km2
limited-access former nuclear production and research facility located in the coastal plain
of South Carolina that is owned and operated by the U.S. Department of Energy (White
and Gaines 2000). The SRS was created to provide nuclear weapons materials at the
beginning of the Cold War in 1951 (Savannah River Nuclear Solutions, LLC 2011) and
now has five decommissioned nuclear reactors, radioactive materials processing
facilities, and retired coal power plants (White and Gaines 2000). The remainder of the
site is dominated by natural habitats and consists of managed pine stands (54%), wetlands
(23%), upland hardwood and mixed forest (11%), grasslands (9%), and upland scrub
forest (3%; White and Gaines 2000; DeVault et al. 2004). Wetlands and other aquatic
habitat on the SRS include bottomland and swamp forests, creeks, streams, two large
cooling reservoirs, upland depressions, and Carolina bays (Lide 1994).
Upon development of the SRS, the conversion of farmland to forest created a
restricted access wildlife refuge currently occupied by 54 species of mammals, 255
species of birds, 100+ species of reptiles and amphibians, and almost 100 species of
freshwater fish (Wike et al. 2006). However, the construction and operation of the
nuclear facilities affected approximately 3,000 hectares of land, created ~2,000 hectares
25
of cooling reservoirs, and all but one stream on site received releases of thermal effluents
(White and Gaines 2000). Contamination from nuclear production, processing, waste
disposal, and power generation is known to exist on the SRS, including radiocesium from
leaking fuel elements and reprocessing operations (White and Gaines 2000). Trace
element contamination also exists on site, primarily in the form of coal fly-ash (Cherry et
al. 1979), and waste mercury from production activities that leaked from seepage basins
(Horton 1974).
The D-Area coal-fired power plants on the SRS were operational from 1953-2012
and during this time deposited sluiced fly ash into a series of basins that drain into Beaver
Dam Creek, a tributary of the Savannah River, and the surrounding wetlands (Gaines et
al. 2002, Bryan et al. 2012, USDOE 2012). Water and biota in the basins and creek
watershed contain elevated levels of aluminum (Al), As, Cd, Cr, iron (Fe), Hg,
manganese (Mn), nickel (Ni), Se, and Zn (Cherry et al. 1979, Rowe et al. 1996) and
numerous studies have documented bioaccumulation and adverse effects from
contaminant exposure in this system for a variety of organisms such as bacteria, aquatic
invertebrates, amphibians, fish, turtles, alligators, and birds (Hopkins et al. 1999, Hopkins
et al. 2000, Rowe et al. 2002, Stepanauskas et al. 2005, Wright et al. 2006; Bryan et al.
2012).
In addition, more than 2.09e+13 Bq of radiocesium was discharged into aquatic
ecosystems on the SRS, representing an important pathway of exposure for numerous
wildlife species (Carlton et al. 1992). In particular, between 1961 and 1964
approximately 5.7 X 1012 Bq of radiocesium was released into Pond B, a reactor cooling
reservoir on the SRS, with documented use by 12 waterfowl species (Ashley and Zeigler
26
1980, Mayer et al. 1986). Past work has quantified levels of radiocesium in Pond B
sediments (Brisbin et al. 1974) and related levels to those found in a variety of animals,
including migratory game birds that can disperse contaminants extensive distances
outside SRS boundaries (Brisbin et al. 1973, Fendley et al. 1977, Brisbin and Vargo
1982, Kennamer et al. 1998).
Collection of Samples
We collected muscle and liver samples from wild pigs and waterfowl/waterbirds
and muscle samples from gray squirrels between late 2012 through early 2015 throughout
various contaminated and uncontaminated areas on the SRS. Sampling locations varied
by species depending upon their abundance and distribution around specific areas of
known SRS operational activities and history of contamination and included the D-Area
ash basins, Fourmile Branch, Pond B, Pond A/R-canal, and various uncontaminated or
generalized areas including L-Lake (Figure 2-1). Additional muscle and liver samples
from wild pigs were collected in Georgia (from Dooly, Macon, Pulaski, Terrel, and
Randolph Counties; courtesy of Jager Pro, Inc.) to serve as reference samples for this
species. We collected squirrels (D-Area ash basins, Fourmile Branch, Pond B/R-canal,
and Tim’s Branch Beaver Pond) and waterfowl/waterbirds (D-Area ash basins, Fourmile
Branch, Pond B/R-canal, Tim’s Branch Beaver Pond, and L-Lake) via shot gun whereas
wild pigs were trapped in box or corral traps throughout the SRS (largely by SRS pig
control contractors) and central GA and euthanized via gunshot to the head. All animal
handling practices and euthanasia were carried out with accompanying federal and state
collecting permits and in accordance with University of Georgia Animal Care and Use
guidelines under protocol A2012 12-010-Y3-A5.
27
Seasonal timing of collections varied by species with most SRS pigs collected in
late summer through fall (GA wild pigs in spring), waterfowl/waterbirds from late fall
through winter, and squirrels in fall. Upon collection, all squirrels and
waterfowl/waterbirds were weighed, sexed, and frozen at -20°C for later dissection. We
collected a muscle sample from the upper hind leg of all collected squirrels.
Waterfowl/waterbirds were whole-body counted for radiocesium and a subset were
dissected to collect breast muscle and liver tissues for additional radiocesium and trace
element analyses to provide data for comparison studies as well as calculations for
consumption limits of muscle tissue. Wild pigs were sexed and weighed in the field
where we also collected tissue samples from the upper hind leg muscle and liver for
contaminant analyses. Wet weights of all muscle and liver samples were recorded before
samples were freeze-dried and weighed again, then homogenized into a powder using a
coffee grinder. Grinder canisters were cleaned with a 5% nitric acid solution and dried
between uses. All wild pig and squirrel samples were tested, along with the subset of
waterfowl for which we collected muscle and liver tissue, for trace elements and
radiocesium concentrations for comparison against published effects levels when
available. We also used these data to calculate human consumption limits for muscle
tissue.
Elemental Analysis
Trace element [Cr, Ni, Cu, Zn, As, Se, Cd, Pb, and uranium (U)] analysis was
conducted on all muscle and liver samples; samples were also analyzed for total mercury
(THg) content. For trace element analysis approximately 250 mg of dry sample was
microwave digested (MARSX Xpress, CEM Corporation, Matthews, NC) with 10.0 ml
28
trace metal-grade nitric acid (70% HNO3). Following digestion, samples were brought to
a final volume of 15.0 ml with Milli-Q (18MΩ) water before analysis by inductively
coupled plasma mass spectroscopy (Nexlon 300X ICP-MS; Perkin Elmer, Norwalk, CT)
according to QA/QC protocols outlined in EPA Method 6020A (USEPA 2007). The
minimum detection limits (ppm) for each element was: Cr (0.54), Ni (0.84), Cu (0.67),
Zn (7.47), As (0.39), Se (0.49), Cd (0.40), Pb (0.36), U (0.44). For quality control
purposes certified reference material (TORT-3 lobster hepatopancreas; National Research
Council, Ottawa, ON, Canada), a blank, and a digestion replicate were run for every 20
samples. Mean percent recoveries ranged from 85%-105% for elements in certified
reference materials and all element concentrations are presented as parts per million
(ppm) on a dry mass basis.
To analyze total mercury (THg), 30-50 mg subsamples of the freeze-
dried/homogenized tissues were analyzed by thermal decomposition, catalytic
conversion, amalgamation, and atomic absorption spectrophotometry (DMA 80;
Milestone, Shelton, CT, USA) according to U.S. EPA) method 7473. The instrument
detection limit (IDL) for this method is 0.01 nanograms (ng) of total mercury. Within
each set of 10 samples we included a replicate, blank, and two standard reference
materials (SRM; TORT-2 lobster hepatopancreas or TORT-3 lobster hepatopancreas, and
PACS-2 marine sediment, National Research Council of Canada, Ottawa, ON) to ensure
quality assurance and solid SRMs were used to calibrate the instrument. Method
detection limits (MDLs; threefold the standard deviation of procedural blanks) averaged
0.0004 ppm dry mass. Mean percent recoveries of THg for the SRMs TORT-2, TORT-3,
29
or PACS-2 were 102.1 ± 1.6, 106.0 ± 3.8, and 102.6 ± 5.0, respectively. Concentrations
are presented as parts per million (ppm) on a dry mass basis.
Radiocesium Analysis
Waterfowl/waterbirds were whole-body counted with a 10·2-cm x 15·2-cm NaI
(Tl) gamma detector (Bicron Model:6H3Q/5; S/N:BJ-124R) coupled to an IBM 300-GL
Personal Computer (Windows 98 OS) containing an onboard Canberra MCA card and
controlled by Canberra Genie 2000 gamma spectroscopy software (Version 1.3; entire
system located in SREL Lab 120). A counting window (Region of Interest-ROI) of 596–
728 kiloelectron-Volts (keV) centered on 662 keV was used to record total detector
absorption events from the radiocesium emission of 662 keV photons. The system was
calibrated daily, as counting took place, with a traceable radiocesium calibration disc
(New England Nuclear Gamma Reference Disc Source Set; Catalogue No. NES-101S;
radiocesium disc; 1.04 microCuries on 10/2/1985) by adjusting the system amplifier gain
control to center the disc-generated peak on channel 331 (661.7keV). Generally, 30-min
count times (1800 sec) were used for counting collected/sacrificed birds (whole-body;
frozen) and backgrounds (empty chamber), while 15-min count times (900 sec) were
used for counting aqueous standards (the ILB Series of standards; containing known
radiocesium quantities [Becquerels (Bq)]; decay corrected to count dates). Background-
corrected count rates (counts per second; cps) from the ILB Series of standards were used
to produce mass-specific count yields which were in turn used to produce a predictive
equation of expected yields from bird mass (in grams; yield=0.4449*mass-0.343;
R2=0.97). Finally, background-adjusted bird count rates and the birds' mass-specific
30
yields were used to determine radiocesium content of birds (whole-body Bq, Bq/g,
pCi/g).
All liver and/or muscle samples were freeze-dried and powdered then packed into
scintillation vials. We analyzed these samples for radiocesium activity using a Packard
Cobra II Auto-Gamma Counter (Model Cobra II 5003) with a single 3-inch through-hole
NaI detector. A counting window (Region of Interest-ROI) of 580–754 kiloelectron-Volts
(keV) centered approximately on 662 keV was used to record total absorption events
from the radiocesium emission of 662 keV photons. The system was auto-calibrated
daily, as counting took place, with a traceable radiocesium source (SREL-0113; 0.1
microCuries on 10/2012). Generally, 60-min count times (3600 sec) were used for
counting dry, powdered samples (packed into tubes) and backgrounds (empty tubes) that
were arranged in every fifth counting position. Four standards were prepared from
commercially-available chicken breast muscle tissue that was dried, homogenized into
powder, spiked with known quantities of radiocesium (745 Bq in each spiked standard;
decay-corrected for the date of preparation), and then loaded/packed into 4 scintillation
tubes in 1-gram increments ranging from 1-4g. Background-corrected count rates (net
counts per second; ncps) recorded for these spiked standards were used to produce
estimates of mass-specific count yields (a ratio of measured net count rate [ncps] to
expected disintegration rates [dps] for the known radiocesium activity) at each of 4
available sample height settings relative to the NaI detector of the Cobra II system.
Sample position #4 produced count yields that varied little across all sample masses (SD
< 0.007) and so samples/backgrounds were all counted using sample height #4 setting,
with an averaged count yield value of 0.2213 used as a constant in radiocesium content
31
determinations for unknowns. Specifically, background-adjusted dry tissue count rates
divided by the yield constant were estimated as the radiocesium content of dry tissue
samples (total Bq, Bq/g [dry mass], and Bq/g [wet mass]). Minimum Detectable
Concentrations (MDCs) were calculated for all radiocesium analyses using the equation
of Lloyd Currie (Currie 1968).
Statistical Analysis
SRS wild pig samples (for trace elements and radiocesium) were compared to
samples collected in central Georgia (see description above); also sex differences
between wild pigs were explored. Trace elements in SRS squirrels were compared
between the D-Area ash basin and all other SRS areas combined due to previous
determination of high concentrations of certain elements (e.g., Se) in that system for
some biota. Radiocesium in squirrel muscle samples was compared between the
radiocesium-contaminated locations combined (Fourmile Branch beaver pond and Pond
B/R-canal) against two separate areas without radiocesium contamination (D-Area ash
basins and Tim’s Branch). Preliminary analyses of trace element concentrations in
waterfowl/waterbirds suggested the D-Area ash basin samples (collections only of diving
ducks and waterbirds) were distinct from the remaining SRS locations. Therefore, we ran
separate analyses on diving and dabbling ducks and for diving ducks compared elemental
concentrations (Cr, Cu, Zn, Se, Hg, Cd) in tissues of individuals collected from the D-
Area ash basins to tissue concentrations in individuals collected from other SRS sites
combined. Concentrations of trace elements found in dabbling ducks and waterbirds are
presented for reference but no statistical comparisons were made between tissues
concentrations because of small sample sizes and that there were not dabbling ducks
32
represented from collections at the D-Area ash basins. Radiocesium in
waterfowl/waterbirds was quantified for all birds but only compared between locations
where we had sufficient numbers of diving ducks: Pond B/R-Canal, Fourmile Beaver
Pond, L-Lake, and D-Area ash basins.
Trace elements for which >50% samples were below detection limits (BDL) were
excluded from analyses; all remaining concentrations BDL were replaced with 50% of
the respective minimum detection limit (MDL; Hall et al. 2009; Fletcher et al. 2014). We
tested all element results distributions for normality (Shapiro-Wilk test p<0.05) in R (R
Core Team 2012) and subsequently log-transformed all data prior to inclusion in analyses
to improve distributions. We used Multiple Analysis of Variance (MANOVA) models
with the Pillai–Bartlett statistic in R (Package stats version 3.3.1) to test whether element
concentrations differed between locations. Muscle and liver samples were tested
separately because trace element concentrations have been shown to differ between these
tissue types in previous toxicology studies (Scheuhammer 1987, Boening 2000, Mason et
al. 2000, Farkas et al. 2003, Ikem et al. 2003, Coğun et al. 2006, Havelková et al. 2008).
Potential correlations between tissues were assessed for each element and species group
(e.g., Se in muscle and liver for SRS wild pigs) and between elements for each tissue
(e.g., Se vs Hg in SRS wild pig muscle samples) with Spearman Rank correlations on
non-transformed data. Spearman’s correlation test ranks the data, controlling for the non-
normal distribution before making comparisons.
Radiocesium concentrations in several muscle and liver samples were below
background levels, resulting in negative values. Negative values and the very small
positive concentrations in samples below the MDCs are often not reported or are set to
33
some arbitrary indicator value, which biases the mean and variance of such data (Gilbert
and Kinnison 1981, Newman et al. 1989). Therefore, we included all negative values, as
well as those below the MDCs, when determining average radiocesium concentrations.
These data were non-normally distributed, therefore, we minimally scaled the data by
adding a number sufficient to remove all negative values and log transformed values
prior to inclusion in models; data were back-transformed and scaling removed before
inclusion in tables. We created separate Analysis of Variance (ANOVA) models for each
taxa of animals and if there were significant differences found in radiocesium
concentrations between study locations we conducted Tukey post-hoc tests to elucidate
differences among sites. We present geometric means and associated standard errors due
to unbalanced sample sizes between groups.
Human Consumption Limits
Based on observed concentrations of key trace elements known to have
deleterious effects on human health (As, Se, Hg; see results), we derived human
consumption limits for wild pigs, squirrels, as well as dabbling and diving ducks
independently. The EPA provides chronic oral reference dose limits for each of these
trace elements; these are levels at which an individual could be exposed daily over the
course of their life and not expect negative health consequences (citation). Human
consumption limits were calculated with both average and maximum concentrations of
the aforementioned trace elements found in muscle samples.
Nearly all Hg in muscle of higher trophic level organisms such as birds and fish is
in the methylmercury (MeHg) form; 80-100% in muscle of piscivorous birds and up to
98% in fish muscle (Wiener et al. 2003; Evers et al. 2005). We tested for total THg,
34
which would show all mercury, but since the majority of the human exposure is to MeHg
instead of elemental mercury and this is the most bioavailable form for humans to absorb,
calculations of consumption limits were done with EPA limits for MeHg (USEPA 1999a,
2001; Hall et al. 2009). MeHg also is the most toxic form of mercury and can cause
severe neurological damage to humans and wildlife (Grandjean et al. 1999, Clarkson et
al. 2003).
The EPA established chronic oral reference doses of 0.0001 milligrams per
kilograms per day (mg/kg/day) for methylmercury, 0.005 mg/kg/day for selenium
(USEPA 1999b) and 0.0003 mg/kg/day for arsenic (USEPA 1998) were used in
calculating consumption limits. These levels are daily dose exposures that alone are
unlikely to produce appreciable deleterious effects over a lifetime of exposure; as
exposures increase above the reference doses so does the risk of adverse health effects.
We used average standard weights for adults and children (70 kg or 154 lbs and 16 kg or
35 lbs, respectively), as well as meal sizes (227g or ½ lb for adults and 113g or ¼ lb for
children) for our calculations based on established EPA methods for fish advisory
calculations.
Since most meat is neither eaten raw (wet mass) or completely devoid of water
(dry mass) we also calculated trace element concentrations based on what the average
person would be exposed to from consuming cooked meat. For wild pig calculations we
utilized the average percent moisture loss during cooking found for squirrel and deer
muscle (17.9%; Holben 2002) and for ducks we utilized the average percent moisture loss
in cooked duck breast muscle (28.2%; Omojola 2007). We used these values to amend
concentrations of trace elements that would be found in cooked muscle for consumption
35
limit calculations. This gives a higher concentration for the trace element than normally
reported wet weight concentrations, but more accurate to real world scenarios for human
consumption of game meat. Mean radiocesium concentrations (Bq/g) in raw muscle
samples for wild pigs, squirrels, and waterfowl/waterbirds that were above MDCs were
compared to the European Economic Community limit of 0.600 Bq/g for fresh meat
(EEC 1986).
RESULTS
For this study 108 wild pigs, 24 squirrels, and 130 waterfowl/waterbirds were
collected; the number of muscle and liver samples tested for trace elements and
radiocesium by taxa and the sex ratios of those groups is presented in Table 2-1.
Wild Pigs
We quantified trace element concentrations in muscle for 88 SRS wild pigs and
further analyzed liver element concentrations for a subset of 30 of these individuals. In
addition, we quantified trace element concentrations in muscle and liver tissues for 20
non-SRS wild pigs (Table 2-2). Concentrations of six trace elements (Cr, Cu, Zn, Se, Hg,
Cd) were found to be above the MDLs in either muscle or liver and were used in
statistical analyses. MANOVA’s testing for differences in trace element concentrations in
muscle and liver between wild pigs captured on the SRS and those in Georgia revealed
some differences between the two locations. Post-hoc analyses indicated that
concentrations of three elements (Hg, Se, and Zn) were higher in SRS samples in one or
both tissues whereas chromium concentrations were higher in GA pigs (GA muscle Cr
= 3.10 ppm vs. SRS muscle Cr = 0.70 ppm, dry mass; Table 2-2).
€
x
€
x
36
Several trace element concentrations were correlated for both SRS and Georgia
samples (Tables 2-3 and 2-4). In particular, for SRS wild pigs, which had higher Se
concentrations than Georgia wild pigs, Se in muscle was positively correlated, though not
strongly, to all the other elements (Table 2-3). For SRS wild pigs, Se and Cd in liver were
moderately and positively correlated (Table 2-3). For Georgia wild pigs, where Cr was
higher than in SRS wild pigs, Cr was positively related to Cu and Zn in muscle (Table 2-
4). In addition, Se and Hg were both positively correlated between muscle and liver
samples for both SRS and Georgia wild pigs (Tables 2-3 and 2-4).
Comparisons of radiocesium tissue concentrations (Table 2-5) revealed SRS
muscle samples were nearly five times higher (geometric = 0.295 Bq/g, dry mass) than
those sampled in Georgia (geometric = 0.057 Bq/g, dry mass). Higher concentrations
were observed in liver samples as well (Tables 2-5 and 2-6). No differences in
radiocesium tissue concentrations were observed between male and female pigs (Table 2-
6).
Squirrels
Twenty four gray squirrels were collected across the SRS. Concentrations of four
trace elements (Cr, Cu, Zn, Hg) were found to be above the MDLs in squirrel muscle and
were used in statistical analyses. Results of our MANOVA analysis comparing element
concentrations between squirrels collected near the D-Area ash basins versus all other
SRS locations failed to detect any significant differences between sites (Table 2-7). No
correlations were found among trace elements in squirrel muscle tissue.
Radiocesium concentrations in squirrel muscle tissue from locations on the SRS
(Table 2-8) suggested among-site differences, with samples from locations having known
€
x
€
x
37
releases of radiocesium tending to have higher concentrations (Table 2-9). Statistical
comparisons with ANOVA indicated that muscle radiocesium concentrations in squirrels
collected from the Fourmile Branch and Pond B/R-Canal area were highest, (geometric
= 1.27 Bq/g, dry mass), and significantly higher than squirrels from either Tim’s
Branch beaver pond (geometric = 0.38 Bq/g, dry mass) or the D-Area ash basins
(geometric = 0.29 Bq/g, dry mass; Table 2-9).
Waterfowl/Waterbirds
Twelve species of waterfowl and three species of waterbirds collected on the SRS
were included in various analyses; muscle and liver samples for 66 and 70
waterfowl/waterbirds, respectively, were tested for trace elements (Table 2-10). Our
MANOVA indicated Se (muscle and liver) and Zn (muscle only) concentrations were
elevated in D-Area ash basin diving ducks compared to diving ducks collected from other
SRS locations, although ducks collected from other SRS locations had higher Cu
concentrations in liver and higher Cr in muscle (Table 2-11). Arsenic is an important
element found in coal ash and of interest for regulation of wastes, remediation efforts,
and human health risk assessment. More than 50% of the diving ducks sampled in this
study had As concentrations in muscle and liver that were BDL. Therefore, we did not
compare As tissue concentrations among sites. However, of the diving ducks collected
from the D-Area ash basins, 29% and 16% of muscle and liver samples, respectively had
levels that were BDL, compared to 91% and 86% from other SRS locations. Here we
report As concentrations in both muscle and liver for diving ducks collected at the ash
basins (muscle = 0.78±0.10, 0.20-2.07 ppm, dry mass; liver = 2.04±0.35, 0.20-6.95
ppm, dry mass) and other SRS locations (muscle = 0.29±0.07, 0.20-2.74 ppm, dry
€
x
€
x
€
x
x x
x
38
mass; liver = 0.42±0.13, 0.20-4.80 ppm, dry mass) for comparison purposes with other
studies.
Dabbling ducks were collected from three sites on the SRS. Mallards (Anas
platyrhynchos) were collected on the Fourmile Branch beaver pond (n=10) and Wood
Ducks (Aix sponsa) were collected from a beaver pond on Tim’s Branch (n=5), Fourmile
Branch beaver pond (n=1), and R-Canal (n=1). Concentrations of those elements (Cr, Cu,
Zn, Se, Hg, Cd) with the majority of samples > MDL are presented in Table 2-12.
Waterbirds analyzed for elemental concentrations included American Coots
(Fulica Americana) (n=4) and Pied-billed Grebes (Podilymbus podiceps) (n=3) from the
D-Area ash basins, and Double-crested Cormorants (Phalacrocorax auritus) (n=4) from
Pond B. Concentrations of elements (Cr, Cu, Zn, As, Se, Hg, Cd) found in these tissue
samples are presented in Table 2-13. The highest Hg level found during this study was in
the liver tissue of a cormorant collected at Pond B (169.4 ppm, dry mass; Table 2-13).
Due to small sample sizes no statistical testing was performed on dabbling ducks or
waterbirds.
Spearman’s rank correlation revealed several trace element concentrations were
correlated in waterfowl/waterbird tissue samples (Table 2-14). Se concentrations in
muscle were positively correlated to Cu, Zn, and Hg, whereas liver Se concentrations
were positively related to Cr, Cd, and Hg (Table 2-14). Cu concentrations also were
positively correlated to Zn in muscle. Only concentrations of Zn were correlated between
muscle and liver, positively, albeit weakly (Table 2-14).
All 130 waterfowl and water birds collected from the SRS were whole-body
counted for radiocesium. Muscle and liver tissues also were counted for a subset of 98 of
x
39
individuals (Table 2-15). Comparisons of whole-body, muscle, and liver tissues of diving
ducks among four SRS locations indicated radiocesium concentrations in individuals
from Pond B were significantly higher (whole-body: geometric = 0.113 Bq/g, fresh
mass; muscle: geometric = 0.163 Bq/g, wet mass; liver: geometric = 0.149 Bq/g,
wet mass; Table 2-16) than concentrations from the three other SRS locations. Sample
concentrations from Fourmile Branch beaver pond, L-Lake, and the D-Area ash basins
were not statistically different from each other.
Human Consumption Limits
After accounting for average moisture loss during cooking based on studies by
Holben (2002) and Omojola (2007), we estimated a retained moisture content of 56.4%
for wild pigs, 58.8% for squirrels, and 43.2% for waterfowl which we utilized in
subsequent calculations to obtain concentrations of As, Se, and Hg in cooked muscle. The
values for each group for were obtained from calculating the actual moisture lost in the
drying process and then adjusting the element concentration for estimated moisture lost
during cooking instead of complete drying. Muscle concentrations of Hg resulted in
consumption limits that allowed fewer meals than As and Se so Hg results are used as an
overall guide to the number of meals allowed.
SRS wild pigs had lower allowable meals than Georgia wild pigs based on muscle
concentrations of trace elements, although allowances were still well above average
numbers consumed by hunters in the region (Smith, unpublished manuscript) based on
average levels observed in our samples. Consumption limits based on average
concentrations of Hg found in SRS wild pigs (muscle = 0.062 ppm, cooked mass)
calculated for human adults were 15.2 (1/2 lb) meals per month, while limits calculated
with the maximum observed Hg concentration (muscle = 0.350 ppm, cooked mass)
x
€
x
€
x
€
x
€
x
40
reduced the limit to 2.7 meals per month (Table 2-17). The meals allowable for children
calculated utilizing the average concentration of Hg were 7.0 (1/4 lb) meals per month,
but only 1.2 meals per month at the maximum concentration (Table 2-17). Element
concentrations in squirrels collected from the D-Area ash basins were not statistically
different from those collected at other sites. Concentrations of trace elements in squirrel
muscle were low such that at the average level of Hg (muscle = 0.017 ppm, cooked
mass) a child could consume ~25 meals per month and an adult ~50 meals before
approaching the EPA’s oral reference dose limit.
We calculated that diving ducks from the D-Area ash basin had the lowest
allowable number of monthly meals of all taxa. Consumption limits based on the average
Hg concentrations in diving duck muscle from the ash basins ( = 0.447 ppm, cooked
mass) were 2.1 (1/2 lb) meals per month for adults and only 1/3 of a meal per month at
the maximum Hg concentration (muscle = 3.153 ppm, cooked mass; Table 2-18). For
children these consumption limits are less, only 1.0 (1/4 lb) meal per month at the
average level and 1/10 of a meal at the maximum Hg concentration (Table 2-18).
Calculations for consumption of sampled dabbling ducks from other locations are also
included for comparison (Table 2-18).
Even though radiocesium concentrations in wild pigs on the SRS ( = 0.459
Bq/g, dry mass) were nearly five times higher than GA pigs, only 21% of the SRS muscle
samples were above their respective MDCs. No wild pig muscle sample exceeded the
European Economic Community (EEC 1986) limit of 0.600 Bq/g for fresh meat, although
the maximum we observed (0.579 Bq/g, wet mass) approached the limit. The maximum
radiocesium concentration in a squirrel muscle sample (from Pond B/R-Canal; 0.595
€
x
€
x
€
x
€
x
41
Bq/g, wet mass) approached the limit as well, but levels in uncontaminated areas (D-Area
ash basins and Tim’s Branch) and known radiocesium contaminated areas (Pond B/R-
canal and Fourmile Branch) only averaged 0.1002 and 0.2829 Bq/g, wet mass,
respectively. In the present study, 19 of 98 (19.4%) waterfowl/waterbirds had muscle
radiocesium concentrations in excess of the EEC limit. All 19 of these birds were
collected from Pond B, an area with known radiocesium contamination, and included 1
ring-necked duck (Aythya collaris), 2 double-crested cormorants, and 16 American coots;
63.3% of waterfowl/water bird muscle samples collected from Pond B exceeded the EEC
limit. The maximum muscle concentration of radiocesium observed in any animal was
2.14 Bq/g, wet mass, in an American coot from Pond B.
DISCUSSION
Here we present trace element and radiocesium concentrations in consumable
tissues of North American game species that occur in both contaminated and
uncontaminated areas on the SRS. However, the SRS is but one of many such
contaminated sites as anthropogenic disturbances are pervasive throughout ecosystems
across the globe (Sanderson et al. 2002, Ellis and Ramankutty 2008). Our results revealed
substantive within and among species variability in tissue contamination, likely reflecting
underlying differences in resource selection, diet, behavior, and distribution. Species
collected at a CCW ash basin, for example, consistently had elevated levels of trace
elements known to occur in coal-fly-ash, particularly Se, supporting prior studies that
demonstrated CCW is an important source of exposure to this element. Similarly, we
often observed elevated concentrations of radiocesium in individuals collected from SRS
sites with known histories of operational releases of radionuclides (e.g., Pond B and
Fourmile Branch).
42
We examined concentrations of trace elements in game animals relative to
concentrations known to impact wildlife health and/or reproduction. Knowledge of
effects levels/concentrations for many elements and/or many game species (particularly
for most non-carnivorous mammals) is lacking, but does exist for those elements of
concern relative to human health (e.g., Hg and Se). These “effects levels” are generally
associated with liver concentrations as liver often accumulates greater concentrations of
contaminants than muscle tissue, thus providing the highest detection probability.
Elemental concentrations for game mammals (i.e. wild pigs and squirrels) were generally
below levels known to affect wildlife health and/or reproduction, and we typically saw
higher trace element burdens in wild pigs. Since wild pigs commonly root within soil for
subsurface food items, exposure to contaminated soil is likely higher than for squirrels
who are liable to have less sustained contact with contaminated soils (Taggert et al.
2009).
Of particular interest are our results for waterfowl as they are highly mobile and
migratory, and thus able to disperse contaminants hundreds or thousands of kilometers
from point sources, increasing the risk of exposure for hunters beyond local communities
surrounding contaminated sites. Trace element concentrations found in diving duck
tissues suggested differences in potential exposure/risk among elements and locations. Se
concentrations in livers of diving ducks from the CCW basins averaged > 35 ppm dry
mass, with a maximum observed concentration of 71.5 ppm; these levels are above
concentrations considered potentially toxic (10-20 ppm dry mass) or toxic (> 20 ppm dry
mass; Ohlendorf and Heinz 2011) to waterfowl, although currently little is known about
the potential health effects to individuals of populations of waterfowl/waterbirds naturally
43
exposed to suites of trace elements. Average Se concentration in diving ducks collected
from other SRS sites with no known releases of Se was only 9.33 ppm. However, four
samples from non-CCW sites were within the potentially toxic (n=3) and toxic (n=1)
ranges. It is unknown whether these individuals acquired high Se burdens from other SRS
locations or off-site areas within their breeding range or along their migratory route.
Interestingly, the highest concentration of Se (83.65 ppm) was observed in the liver of a
Ring-necked duck collected from Pond B. Ring-necked ducks were infrequently sighted
at the ash basins in our study, suggesting this individual likely accumulated Se elsewhere
within its breeding or migratory range. Liver Se levels in dabbling ducks (mallards and
wood ducks) from areas with no known Se contamination did not exceed 1.5 ppm,
although Se in livers of the waterbird species averaged near or above the level considered
toxic. Given the high trace element concentrations we observed in waterfowl collected
from the ash basin site, and that there are currently 735 similar surface impoundments for
CCW in the U.S. (and many more worldwide), additional studies of a broader scope are
needed to more clearly elucidate potential risks to both wildlife and waterfowl hunters.
Concentrations of Hg in avian liver tissue were generally below levels of concern,
although liver Hg concentrations for a small number of diving ducks from the ash basins
(n=2) and non-ash basin sites (n=1) exceeded levels of concern (2 ppm wet mass, ~8-10
ppm dry mass) thought to result in adverse health effects (Shore et al. 2011). All dabbling
ducks, coots, and grebes had liver Hg concentrations below the level of concern.
However, highly piscivorous cormorants from Pond B had the highest average and
maximum Hg liver concentrations (~60 ppm and ~169 ppm Hg dry mass, respectively) of
44
any animal analyzed in this study, levels that greatly exceeded the threshold at which
adverse health effects could be expected.
Across all taxa surveyed, our results suggest that for humans Hg was the element
that produced the most restrictive allowance of meals per month before exceeding EPA
chronic oral exposure limits. At average Hg concentrations for wild pigs, a relatively high
number of meals for adults and children (15.2 and 7.0 respectively) would be permitted,
although when the maximum concentration was considered allowances dropped to 2.7
meals/month for adults and 1.2 for children, consumption rates relevant to average levels
reported for hunters in the southeastern U.S. (Smith, unpublished manuscript). GA wild
pigs had meal allowances that greatly exceeded those from the SRS (51.6 and 23.7 meals
per month for adults and children, respectively at average Hg levels in muscle). At
maximum concentrations meal allowances of 30.0 and 13.7 meals per month were
calculated for adults and children, respectively. Foraging habits of squirrels appeared to
limit contaminant uptake at sites with higher known concentrations of trace elements in
sediments (i.e. D-Area and Fourmile Branch) and thus consumption of squirrels appear to
be less of a risk for human exposure as evidenced by the large number of meals allowed
(50+ for adults and 25+ for children) when compared to wild pigs and waterfowl. In
contrast to the liberal meal allowances for wild pigs and squirrels even from areas with
known contaminant releases, consumption limits of diving ducks for adults and children
based on average Hg concentrations from the ash basin were 2.1 and 1.0 meal(s) per
month respectively. This was further reduced to 0.3 meals for adults and 0.1 meals for
children when considering maximum observed concentrations in muscle tissue. These
data suggest waterfowl use of areas contaminated with CCW pose a potential risk to not
45
only the waterfowl themselves, but also hunters and their families, especially pregnant
mothers and children (Duchesne et al. 2004). Use of CCW impoundments by waterfowl
as well as uptake rates of contaminants for waterfowl that use contaminated
impoundments within either their breeding or migratory ranges are largely unknown, but
represent an important area of future study.
Unfortunately, as with certain trace elements, “effects” levels or concentrations of
radiocesium relative to wildlife health is lacking (Hinton 1998). Thus, as a conservative
approach, we discuss the observed concentrations and trends relative to human
consumption levels of concern established by the European Economic Community for
radiocesium in fresh meat. Unlike other elements, radiocesium generally accumulates to
higher concentrations in muscle tissue than in liver tissue (e.g., Brisbin and Smith 1975,
Potter et al. 1989), as was observed for wild pigs and waterfowl in our study.
Radiocesium concentrations in wild pigs on the SRS were about five times higher than in
pigs collected from central Georgia. Still, only 21% of the SRS pig muscle samples were
above their respective MDCs. No wild pig muscle sample exceeded the EEC limit for
fresh meat and thus currently consumption of wild pigs from the SRS does not appear to
be a danger to human health. However, pig behavior may increase exposure to and
subsequent accumulation of numerous contaminants including radiocesium; thus in areas
of suspected contamination wild pigs that may be taken as game meat should be tested.
Levels of radiocesium in squirrel muscle tissue from areas with known
radiocesium releases (i.e., Fourmile Branch and Pond B/R-Canal) were higher than in
squirrels from locations with no such known contaminant releases (i.e., D-Area ash
basins and Tim’s Branch beaver pond). The maximum radiocesium concentration in a
46
squirrel muscle sample (from Pond B/R-Canal; 0.595 Bq/g, wet mass) approached the
EEC (1986) limit of 0.600 Bq/g in fresh meat. Squirrels have smaller home ranges than
larger or more mobile species, like wild pigs and waterfowl, and therefore are more
localized threats when considering contamination sources.
Radiocesium concentrations were higher in diving ducks from Pond B (whole-
body geometric = 0.113 Bq/g, fresh mass) than other areas of SRS. However, as
previously reported by Brisbin et al. (1973), American coots had the highest radiocesium
levels among the waterfowl species surveyed. In our study, 19 of 98 (19.4%)
waterfowl/waterbirds sampled had muscle radiocesium concentrations in excess of the
0.600 Bq/g, fresh mass, EEC (1986) limit, all from Pond B; thus 63.3% of the Pond B
birds exceeded the consumption limit. It should be noted that cormorants are not eaten
and coots are less popularly consumed than many other species of waterfowl/waterbirds.
Whole-body radiocesium concentrations in American coots on Pond B ( = 0.94 Bq/g,
fresh mass) were about 7 times higher than Pond B diving ducks. These differences
between coot and diving duck radiocesium levels have been attributed to species
differences in diet (e.g., Brisbin et al. 1973). However, species differences in residence
time on contaminated sites, which is not well understood at this point, has the potential to
explain substantial variation as well. Additional studies are needed to elucidate the effects
of residence time on radiocesium accumulation for a greater diversity of
waterfowl/waterbirds and other wildlife known to utilize radionuclide-contaminated sites
worldwide.
€
x
€
x
47
MANAGEMENT IMPLICATIONS
Our results revealed differences in contaminant concentrations among species,
tissue types, elements, and sampling locations. Although our sampling occurred on the
SRS, a location with known contamination, the majority of tissue samples analyzed were
below levels known to adversely affect wildlife health and in the case of radiocesium fell
below established EEC limits for human consumption. However, some samples collected
from locations with known elevated levels did exceed wildlife health thresholds or EEC
limits. These data fill critical knowledge gaps regarding trace element and radiocesium
accumulation in multiple game species utilizing contaminated areas.
While we continue to deal with environmental contamination from recent
catastrophes like the nuclear accident at Fukushima and the failed ash basin
impoundment that sent coal-fly-ash into the Emory-Clinc-Tennessee River System, we
need to consider how these and future environmental practices affect wildlife that can
expose humans hunters and their families to contaminants. The research presented herein
represents an important advance in our understanding of potential exposure levels of
wildlife inhabiting contaminated areas, however, long-term monitoring of contaminant
exposure in game species and the subsequent effects on resident or migrant wildlife and
human hunters across larger spatial scales is essential. Species with large home ranges or
that are migratory should be a priority for monitoring as they present risks to the public
far from contaminant sources (Taggart et al. 2011). Also most often we analyze human
consumption risks on an individual species basis; but we do know that most hunters
consume multiple species so future studies should treat consumption limit calculations
more holistically, incorporating fish, as well as multiple game species.
48
Table 2-1. Sample sizes of game species collected from the Savannah River Site (SRS)
and analyzed for trace elements and radiocesium in 2012-2015; sex ratios are also
indicated.
Species Sampled Analyses Muscle M/F Liver M/F SRS Wild pigsa Trace Elements 88 (50/37) 30 (14/15)
Radiocesium 82 (49/33) 26 (12/14) GA Wild pigs Trace Elements 20 (12/8) 20 (12/8)
Radiocesium 19 (11/8) 20 (12/8) SRS Squirrels Trace Elements 24 (9/15) - -
Radiocesium 24 (9/15) - - SRS Waterfowlb Trace Elements 66 (41/25) 70 (45/25) Radiocesium 98 (60/38) 98 (60/38)
aSex was not determined for one SRS wild pig.
b130 total waterfowl were whole-body counted for radiocesium.
49
Table 2-2. Comparisons of trace element concentrations (ppm, dry mass) in muscle and liver tissues of wild pigs collected
from the Savannah River Site (SRS) and those collected from five counties in Georgia (GA) in 2012-2015. See Table 2-1 for
sample sizes of muscle and liver tissues analyzed.
!! ______SRS Wild Pigs___ ______GA Wild Pigs___ Elementa Tissue Mean SE Range Mean SE Range F Pb Cr Muscle 0.70 0.04 0.27-1.62
3.10 0.65 0.76-10.82
57.25 <0.0001
Liver 0.70 0.11 0.27-3.02
1.58 0.21 0.77-4.02
16.63 <0.001
Cu Muscle 9.87 1.42 2.85-84.60
7.03 1.28 3.24-29.98
0.87 0.35
Liver 24.30 3.96 0.34-89.99
19.94 2.71 10.08-63.64
0.67 0.42
Zn Muscle 86.38 3.38 42.56-211.91
69.08 3.95 40.38-106.15
5.53 <0.05
Liver 113.83 7.31 3.74-203.74
112.51 4.58 90.74-170.66
0.02 0.89
Se Muscle 1.79 0.07 0.25-3.56
0.96 0.12 0.25-1.99
27.27 <0.0001
Liver 2.95 0.16 1.37-4.81
2.21 0.21 1.11-4.38
7.75 <0.01
Hg Muscle 0.14 0.01 0.01-0.80
0.04 0.00 0.02-0.07
21.06 <0.0001
Liver 0.35 0.04 0.06-0.87
0.07 0.01 0.02-0.15
36.44 <0.0001
Cdc Liver 1.14 0.24 0.20-5.22 0.44 0.09 0.20-1.83 ― ― aOver 50% of the following elements were below method detection limits and are not presented in this table: As, Ni, Pb, U,
and Cd (muscle).
bNS=not significant (P > 0.05).
cNot compared statistically due to a low percentage (~45%) of above detection concentrations for the GA wild pigs.
50
Table 2-3. Correlationsa among trace element concentrations in muscle (above diagonal) and liver (below diagonal) tissues of wild
pigs collected from the Savannah River Site (SRS) in 2012-2015 (n=88 muscle, n=30 liver). Only those elements with more than 50%
of the values above detectable limits are included. Correlations between muscle and liver samples for individual elements are
presented on the diagonal in bold.
Cr Cu Zn Se Hg Cd Cr NS 0.29 (0.006) NS 0.21 (0.05) NS ― Cu NS NS NS 0.26 (0.02) NS ― Zn NS NS NS 0.37 (0.0004) NS ― Se -0.50 (0.005) NS NS 0.62 (0.0004) 0.25 (0.02) ― Hg NS NS NS NS 0.45 (0.01) ― Cd -0.43 (0.02) NS NS 0.76 (<0.0001) NS ―
aSpearman correlation coefficients and P values in parentheses; NS=not significant (P > 0.05).
51
Table 2-4. Correlationsa among trace element concentrations in muscle (above diagonal) and liver (below diagonal) tissues of wild
pigs collected from five counties in Georgia (GA) in 2012-2015 (n=20 muscle, n=20 liver). Only those elements with more than 50%
of the values above detectable limits are included. Correlations between muscle and liver samples for individual elements are
presented on the diagonal in bold.
Cr Cu Zn Se Hg Cd
Cr NS 0.59 (0.007) 0.50 (0.03) NS NS ―
Cu 0.48 (0.03) NS NS NS NS ―
Zn NS NS NS NS NS ―
Se NS NS NS 0.84 (<0.0001) NS ―
Hg NS NS NS 0.66 (0.002) 0.50 (0.02) ―
Cd NS NS NS NS NS ― aSpearman correlation coefficients and P values in parentheses; NS=not significant (P > 0.05).
52
Table 2-5. Descriptive statistics for radiocesium concentrations in muscle and liver tissues of wild pigs collected from the
Savannah River Site (SRS) and from five counties in Georgia (GA) in 2012-2015.
Group Compartment Variable N % below
MDCa Mean SE Min Max Median CV (%) SRS Pigs Muscle Bq/g; dry mass 82 49 0.289 0.034 -0.053 2.211 0.209 105.2
Bq/g; wet mass
0.075 0.009 -0.014 0.579 0.054 107.9
MDC; dry mass
0.207 0.010 0.128 0.795 0.185 44.1
wet:dry ratio 76
3.905 0.045 2.218 4.676 3.905 10.5
GA Pigs Muscle Bq/g; dry mass 19 100 0.051 0.013 -0.062 0.151 0.044 113.5
Bq/g; wet mass
0.020 0.006 -0.028 -0.070 0.016 123.6
MDC; dry mass
0.202 0.019 0.128 0.431 0.176 40.2
wet:dry ratio 17
2.585 0.058 2.124 2.955 2.585 9.9
SRS Pigs Liver Bq/g; dry mass 26 85 0.085 0.017 -0.031 0.234 0.081 101.8
Bq/g; wet mass
0.026 0.005 -0.010 0.073 0.023 103.0
MDC; dry mass
0.198 0.014 0.121 0.405 0.168 36.4
wet:dry ratio 18
3.278 0.056 2.805 3.869 3.278 8.7
GA Pigs Liver Bq/g; dry mass 20 100 -0.026 0.012 -0.147 0.097 -0.029 -201.8
Bq/g; wet mass
-0.012 0.005 -0.067 0.046 -0.012 -208.6
MDC; dry mass
0.227 0.013 0.122 0.320 0.235 26.6
wet:dry ratio 19 2.363 0.056 2.026 3.080 2.338 10.6 aMDC=minimum detectable concentration; calculated from Currie (1968).
53
Table 2-6. Radiocesium concentrations (Bq/g, dry mass) in muscle and liver tissues of
wild pigs collected from the Savannah River Site (SRS) and from five counties in
Georgia (GA) from 2012-2015a.
Tissue Location Geometric Mean
Lower 95% CI
Upper 95% CI
Muscle SRS Pigs 0.295 0.222 0.384 GA Pigs 0.057 0.006 0.136 Liver SRS Pigs 0.097 0.042 0.167 GA Pigs -0.048 -0.076 -0.010 Muscle F (N=41) 0.149 0.083 0.241 M (N=60) 0.147 0.088 0.225 Liver F (N=22) 0.040 -0.008 0.101 M (N=24) -0.014 -0.048 0.028
aRadiocesium concentrations differed by location for muscle (ANOVA: F(1,98) = 16.023,
P = 0.0001) and liver (ANOVA: F(1,43) = 22.329, P = <0.0001) tissues; radiocesium
levels did not differ by sex.
54
Table 2-7. Comparisons of trace element concentrations (ppm, dry mass) in muscle of squirrels collected near the D-Area ash
basins and squirrels collected from all other locations on the Savannah River Site (SRS) in 2012-2015.
Ash Basin Squirrels (N=9) Other Squirrels (N=15) Elementa Tissue Mean SE Range Mean SE Range F Pb Cr Muscle 1.20 0.15 0.55-1.86
1.18 0.16 0.27-2.42
0.01 0.94
Cu Muscle 5.21 0.64 3.55-9.84
4.93 0.44 3.20-10.15
0.14 0.71
Zn Muscle 35.94 1.47 31.86-45.57
35.34 1.28 27.78-43.24
0.09 0.77
Hg Muscle 0.04 0.01 0.02-0.10 0.05 0.01 0.01-0.17 0.39 0.54 aOver 50% of the following elements were below method detection limits and were not presented in this table: As, Cd, Ni, Pb,
Se, and U.
bNS=not significant (P > 0.05).
55
Table 2-8. Descriptive statistics for radiocesium concentrations in muscle tissue of squirrels collected from the Savannah
River Site (SRS) in 2012-2015.
Tissue Squirrel Location Variable N
% below MDCa Mean SE Min Max Median CV (%)
Muscle D-Area Ash Basins Bq/g; dry mass 9 56 0.275 0.035 0.163 0.439 0.258 37.9
Bq/g; wet mass 9
0.117 0.015 0.073 0.199 0.098 39.7
MDC; dry mass 9
0.268 0.023 0.186 0.343 0.301 25.8
wet:dry ratio 5
2.370 0.104 1.959 3.112 2.370 13.2
Four-mile/Pond B/R-canal Bq/g; dry mass 9 44 1.159 0.290 0.202 2.181 1.178 75.1
Bq/g; wet mass 9
0.283 0.068 0.059 0.595 0.284 72.6
MDC; dry mass 9
0.188 0.004 0.170 0.203 0.195 6.3
wet:dry ratio 9
2.521 0.181 2.873 4.582 3.874 14.0
Tim's Branch Beaver Pond Bq/g; dry mass 6 17 0.281 0.045 0.096 0.404 0.307 39.4
Bq/g; wet mass 6
0.075 0.014 0.023 0.122 0.078 45.7
MDC; dry mass 6
0.202 0.019 0.159 0.290 0.191 23.2
wet:dry ratio 6 3.878 0.208 2.930 4.310 4.074 13.1 aMDC=minimum detectable concentration; calculated from Currie (1968).
56
Table 2-9. Comparison of radiocesium concentrations (Bq/g, dry mass) in muscle tissue
of squirrels collected from four different locations on the SRS in 2012-2015.a
Location Geometric Mean
Upper 95% CI
Lower 95% CI
D-Area Ash BasinsA 0.292 0.524 0.141
Fourmile/Pond B/R-canalB 1.269 2.027 0.776
Tim's Branch Beaver PondA 0.383 0.746 0.168 aRadiocesium concentrations differed among locations for muscle (ANOVA: F(2,21) =
9.088, P < 0.001). Post-hoc Tukey HSD tests were used for determining pairwise
locational differences. Locations with the same capital letter were not significantly
different (P > 0.05).
57
Table 2-10. Waterfowl and waterbird species collected on the Savannah River Site (SRS) in 2012-2015, with the scientific
names, alpha codes, guild groupings, and sample sizes in various analyses.
Common Name Scientific Name Alpha Code Guild Group Muscle
Elements Liver
Elements Muscle /
Liver 137Cs Whole-body
137Cs
Wood Duck Aix sponsa WODU Dabbling Duck 7 7 10 10 Green-winged Teal Anas crecca GWTE Dabbling Duck 0 0 0 3 Mallard Anas platyrhynchos MALL Dabbling Duck 10 10 10 14 Gadwall Anas strepera GADW Dabbling Duck 0 0 0 2 American Wigeon Anas americana AMWI Dabbling Duck 0 0 0 2 Northern Pintail Anas acuta NOPI Dabbling Duck 0 0 0 1
Canvasback Aythya valisineria CANV Diving Duck 1 1 1 1 Ring-necked Duck Aythya collaris RNDU Diving Duck 8 8 14 23 Lesser Scaup Aythya affinis LESC Diving Duck 13 13 13 14 Bufflehead Bucephala albeola BUFF Diving Duck 8 8 8 9 Hooded Merganser Lophodytes cucullatus HOME Diving Duck 5 5 5 5 Ruddy Duck Oxyura jamaicensis RUDU Diving Duck 7 7 7 9 American Coot Fulica americana AMCO Rail 4 4 23 29
Pied-billed Grebe Podilymbus podiceps PBGR Diver 3 3 3 4 Double-crested Cormorant Phalacrocorax auritus DDCO Diver 0 4 4 4
TOTALS 66 70 98 130 aResidence status on the SRS: R=resident, on the SRS in all seasons; M=migrant, only on the SRS in the fall/winter.
58
Table 2-11. Comparisons of trace element concentrations (ppm, dry mass) in muscle and liver tissues of diving ducks collected
from D-Area ash basins and diving ducks collected from other water bodies on the Savannah River Site (SRS) in 2012-2015.
Ash Basin Divers (n=24) Other Divers (n=18) Elementa Tissue Mean SE Range Mean SE Range F Pb Cr Muscle 0.58 0.06 0.27-1.09
0.94 0.23 0.27-4.05 4.69 <0.01
Liver 1.13 0.11 0.27-2.75
1.18 0.35 0.27-6.53 0.03 0.87
Cu Muscle 34.04 2.87 18.85-78.06
29.62 1.75 7.01-40.05 2.99 0.09
Liver 61.50 10.04 12.57-244.92
226.07 48.51 26.30-699.14 14.37 <0.001
Zn Muscle 39.23 2.10 25.18-64.00
31.12 1.77 8.01-43.72 7.94 <0.01
Liver 118.59 7.03 68.62-211.01
141.25 15.11 66.10-308.93 2.18 0.1500
Se Muscle 12.90 1.29 2.52-23.45
4.40 2.25 0.53-42.04 12.06 <0.001
Liver 35.58 3.65 5.05-71.48
9.33 4.42 2.03-83.65 21.25 <0.0001
Hg Muscle 0.79 0.24 0.07-5.55
0.37 0.11 0.02-1.39 2.05 0.16
Liver 2.67 0.64 0.23-11.50
1.52 0.61 0.09-10.89 1.60 0.21
Cd Liver 2.73 0.69 0.49-13.06 2.37 0.95 0.20-17.18 0.09 0.76 aOver 50% of the following elements were below method detection limits and were not presented in this table: As, Ni, Pb, U,
and Cd (muscle).
bNS=not significant (P > 0.05).
59
Table 2-12. Concentrations of trace elements (ppm, dry mass) in muscle and liver tissues of dabbling ducks collected from the
Savannah River Site (SRS) in 2012-2015.
Elementa Tissue Mallard (n=10)b Wood Duck (n=7)c Mean SE Range Mean SE Range
Cr Muscle 0.88 0.19 0.27-2.28 0.75 0.18 0.27-1.40
Liver 0.69 0.16 0.27-1.66 - - Cu Muscle 19.51 1.51 14.43-31.13 16.57 1.04 13.33-20.31
Liver 110.01 14.44 51.32-176.06 28.55 3.42 21.04-42.26
Zn Muscle 33.46 1.24 27.81-41.73 29.39 2.43 21.91-41.33
Liver 120.07 10.12 84.01-181.69 150.95 14.85 87.26-203.23
Se Muscle 1.57 0.24 0.57-3.13 1.04 0.2 0.24-1.97
Liver 3.84 0.16 2.96-4.63 2.72 0.42 1.54-4.81
Hg Muscle 0.09 0.02 0.04-0.23 0.04 0.01 0.02-0.10
Liver 0.34 0.09 0.13-1.05 0.13 0.04 0.05-0.33
Cd Liver 1.14 0.28 0.20-3.00 0.9 0.23 0.20-2.10 aOver 50% of the following elements were below method detection limits and were not presented in this table: Cr (liver of Wood
Ducks), Cd (muscle of both species), As, Ni, Pb, U.
bAll 10 Mallards were collected at Fourmile Branch beaver pond.
cWood Ducks were collected at Tim's Branch beaver pond (n=5), Fourmile Branch beaver pond (n=1), and R-Canal (n=1).
60
Table 2-13. Concentrations of trace elements (ppm, dry mass) in muscle and liver tissues of other water birds collected from
the Savannah River Site (SRS) in 2012-2015.
Elementa Tissue American Coot (n=4)b Pied-billed Grebe (n=3)c Double-crested Cormorant (n=4)d
Mean SE Range Mean SE Range Mean SE Range Cr Muscle 0.92 0.2 0.55-1.48 0.84 0.09 0.69-1.02 - - -
Liver 1.29 0.18 0.84-1.71 1.35 0.1 1.15-1.49 1.28 0.11 1.05-1.53
Cu Muscle 63.56 3.83 52.79-70.90 37.74 5.91 28.35-48.65 - - -
Liver 41.75 22.09 10.39-105.38 26.91 8.8 14.61-43.96 18.72 1.53 16.21-22.46
Zn Muscle 60.3 3.37 54.81-70.13 56.11 2.58 50.98-59.04 - - -
Liver 114.56 36.09 58.84-216.90 105.95 8.28 94.98-122.17 74.86 2.78 69.96-82.49
As Muscle 0.85 0.27 0.20-1.50 - - - - - -
Liver 2.04 1.32 0.20-5.89 0.37 0.09 0.20-0.48 - - -
Se Muscle 15.33 2.32 10.44-21.59 28.12 4.51 19.18-33.52 - - -
Liver 17.41 3.7 9.72-25.49 41.79 8.62 24.71-52.35 19.52 12.06 3.71-55.39
Hg Muscle 0.06 0.01 0.03-0,08 0.72 0.08 0.61-0.87 - - -
Liver 0.37 0.07 0.24-0.51 1.33 0.2 0.97-1.66 60.64 37.19 5.75-169.39
Cd Liver 0.36 0.09 0.20-0.56 2.49 1.63 0.79-5.75 0.65 0.09 0.47-0.81
a >50% of elements were BDL thus not presented: Cd (muscle all species), As (grebe muscle and cormorant liver), Ni, Pb, U.
bAll 4 coots were collected from D-Area ash basins.
cAll 3 grebes were collected from D-Area ash basins.
dAll 4 cormorants were collected from Pond B.
61
Table 2-14. Correlationsa among trace element concentrations in muscle (above diagonal) and liver (below diagonal) tissues of all
waterfowl/waterbirds collected from the Savannah River Site (SRS) in 2012-2015 (n=66 muscle, n=70 liver). Only those elements
with more than 50% of the values above detectable limits are included. Correlations between muscle and liver samples for individual
elements are presented on the diagonal in bold.
Cr Cu Zn Se Hg Cd Cr NS NS NS NS NS ―
Cu NS NS 0.63 (<0.0001) 0.53 (<0.0001) NS ―
Zn -0.27 (0.02) 0.52 (<0.0001) 0.29 (0.02) 0.52 (<0.0001) NS ―
Se 0.47 (<0.0001) NS NS NS 0.57 (<0.0001) ―
Hg 0.31 (0.009) -0.29 (0.01) -0.37 (0.002) 0.62 (<0.0001) NS ―
Cd NS NS 0.26 (0.03) 0.33 (0.006) NS ― aGiven are Spearman correlation coefficients and P values in parentheses; NS=not significant (P > 0.05).
62
Table 2-15. Descriptive statistics for radiocesium concentrations measured in waterfowl collected from the Savannah River
Site (SRS) in 2012-2015.a
Compartment Variable N % below MDCb Mean SE Min Max Median CV (%)
Whole-body Bq/g; fresh mass 130 56 0.159 0.029 -0.005 1.683 0.01 207.7
! MDCb; fresh mass 130 ! 0.0057 0.0001 0.0023 0.0103 0.0056 26.7
! ! ! ! ! ! ! ! ! !Muscle Bq/g; dry mass 98 62 1.228 0.224 -0.083 8.027 0.1 180.3
! Bq/g; wet mass 98 ! 0.328 0.059 -0.024 2.145 0.028 177.8
! MDC; dry mass 98 ! 0.163 0.002 0.123 0.223 0.161 12.9
! wet:dry ratio 87 ! 3.603 0.021 3.269 4.397 3.581 5.3
! ! ! ! ! ! ! ! ! !Liver Bq/g; dry mass 98 66 0.797 0.151 -0.087 7.566 0.058 187.8
! Bq/g; wet mass 98 ! 0.225 0.042 -0.026 2.07 0.016 186.7
! MDC; dry mass 98 ! 0.152 0.003 0.11 0.281 0.15 17.8 wet:dry ratio 87 3.539 0.029 2.829 4.411 3.514 7.6
aIncludes all waterfowl collected from 6 locations on the SRS: D-Area ash basins, Tim’s Branch beaver pond, L-Lake,
Fourmile Branch beaver pond, R-Canal/Pond A, and Pond B.
bMDC=minimum detectable concentration; calculated from Currie (1968).
63
Table 2-16. Comparisons of whole-body and tissue radiocesium concentrations (Bq/g, wet mass) in diving ducks (Bufflehead,
Canvasback, Lesser Scaup, Ring-necked Duck and Ruddy Duck) collected from four different locations of the SRS in 2012-
2015.a
Compartment Location Geometric Mean
Lower 95% CI
Upper 95% CI Max
Whole-body (N=55) Pond BA 0.1127 0.0693 0.1815 0.2651
Fourmile Beaver PondB 0.0063 0.0016 0.0143 0.0391
L-LakeB 0.0005 -0.0012 0.0031 0.012
D-Area Ash BasinsB -0.0002 -0.0016 0.0018 0.0389
Muscle (N=42) Pond BA 0.1634 0.0663 0.3689 0.6362
Fourmile Beaver PondB 0.0134 -0.0061 0.0584 0.0935
L-LakeB -0.0029 -0.0147 0.0311 0.0204
D-Area Ash BasinsB -0.0091 -0.0141 -0.0005 0.0764
Liver (N=42) Pond BA 0.1485 0.0818 0.2565 0.3036
L-LakeB 0.0047 -0.0105 0.0347 0.0341
Fourmile Beaver PondB -0.0009 -0.0113 0.017 0.0281 D-Area Ash BasinsB -0.0027 -0.0096 0.007 0.0551
aANOVA tests were conducted separately by compartment for a location effect on log-transformed (and scaled) radiocesium
levels (Whole-body: F(3,51) = 45.62, P < 0.0001; Muscle: F(3,38) = 11.57, P < 0.0001; Liver: F(3,38) = 16.19, P < 0.0001), and
followed by post-hoc Tukey HSD tests for determining pairwise locational differences. Locations within compartments with
the same capital letter were not significantly different (P > 0.05). Geometric means are back-transformations of least-squares
means of log-transformed (with scaling removed) radiocesium concentrations.
64
Table 2-17. Monthly allowances of ½ lb. meals for adults and ¼ lb. for children before exceeding the EPA’s oral reference dose
ratings for selenium (Se) and mercury (Hg) for muscle tissue of wild pigs collected from the Savannah River Site (SRS) (n=88) and
five counties in Georgia (GA) (n=20) 2012-2015. Consumption limits based on average concentrations are presented with
consumption limits based on the maximum concentration found in an individual in parentheses. Levels of As were all BDL so
consumption limits are not included.
SRS Wild Pigs GA Wild Pigs Element Age Dry Cooked Raw Dry Cooked Raw Se Adult 26.3 (13.2) 60.2 (30.2) 102.3 (51.2)
48.6 (23.6) 111.5 (54.1) 189.3 (91.9)
Child 12.1 (6.1) 27.7 (13.9) 47.0 (23.5)
22.3 (10.8) 51.2 (24.9) 86.9 (42.2)
Hg Adult 6.6 (1.2) 15.2 (2.7) 25.8 (4.5)
22.5 (13.1) 51.6 (30.0) 87.6 (50.9) Child 3.0 (0.5) 7.0 (1.2) 11.8 (2.1) 10.3 (6.0) 23.7 (13.7) 40.2 (23.4)
65
Table 2-18. The monthly allowances of ½ lb. meals for adults and ¼ lb. for children before exceeding the EPA’s chronic oral
reference dose limits for arsenic (As), selenium (Se), and mercury (Hg) for muscle tissue of diving ducks collected from the D-
Area ash basins (n=24) and other water bodies (n=18) on the Savannah River Site (SRS) 2012-2015. Consumption limits based
on average concentrations are presented with limits based on the maximum concentration found in an individual for each trace
element in parentheses. Levels of As for Dabbling ducks were all BDL so consumption limits are not included.
Ash Basin Diving Ducks Other SRS Diving Ducks Dabbling Ducks Element Age Dry Cooked Raw Dry Cooked Raw Dry Cooked Raw As Adult 3.4 (1.4) 5.9 (2.4) 11.8 (4.8) 9.6 (1.0) 16.9 (1.8) 33.5 (3.6) ― ― ―
Child 1.5 (0.6) 2.7 (1.1) 5.4 (2.2) 4.4 (0.5) 7.8 (0.8) 15.4 (1.6) ― ― ―
Se Adult 3.6 (2.0) 6.4 (3.5) 12.2 (7.0) 16.1 (1.1) 28.3 (2.0) 56.2 (3.9) 34.7 (15.0) 61.1 (26.4) 121.3 (52.4)
Child 1.7 (0.9) 2.9 (1.6) 5.8 (3.2) 7.4 (0.5) 13.0 (1.0) 25.8 (1.8) 15.9 (6.9) 28.1 (12.1) 55.7 (24.1)
Hg Adult 1.2 (0.2) 2.1 (0.3) 4.2 (0.6) 4.2 (0.7) 7.3 (1.2) 15.0 (2.4) 13.0 (4.0) 23.0 (7.1) 45.6 (14.1) Child 0.6 (0.1) 1.0 (0.1) 1.9 (0.3) 2.0 (0.3) 3.4 (0.6) 6.7 (1.1) 6.0 (1.9) 10.5 (3.3) 20.9 (6.5)
66
Figure 2-1. Savannah River Site (SRS) locations targeted for sample collections (wild
pigs, squirrels, waterfowl/waterbirds) for trace elements and radiocesium quantification
in 2012-2015, included the D-Area ash basins, Fourmile Branch, Tim’s Branch, Pond
A/R-Canal, Pond B, and L-Lake.
67
LITERATURE CITED
Ashley, C., and C. Zeigler. 1980. Releases of radioactivity at the Savannah River Plant,
1954 through 1978. Report DPSU 75-25-1. E.I. DuPont deNemours, Savannah
River Laboratory, SC.
Bagshaw C., and I. L. Brisbin Jr. 1984. Long-term declines in radiocesium of two
sympatric snake populations. Journal of Applied Ecology 21:407-413.
Beck, M. L., W.A. Hopkins, and B. P. Jackson. 2013. Spatial and temporal variation in
the diet of tree swallows: implications for trace-element exposure after habitat
remediation. Archives of Environmental Contamination and Toxicology 65:575-
587.
Beck, M. L., W. A. Hopkins, B. P. Jackson, and D. M. Hawley. 2015. The effects of a
remediated fly ash spill and weather conditions on reproductive success and
offspring development in tree swallows. Environmental Monitoring and
Assessment 187:1-25.
Bennett, B. G. 1995. Exposures from worldwide releases of radionuclides.
Environmental Impact of Radioactive Releases. International Atomic Energy
Agency (IAEA), Vienna, Austria, pp. 3-12.
Boening, D. W. 2000. Ecological effects, transport, and fate of mercury: A general
review. Chemosphere 40:1335-1351.
Braune, B. M., and B. J. Malone. 2006. Organochlorines and trace elements in upland
game birds harvested in Canada. Science of the Total Environment, 363:60-69.
Brisbin Jr., I. L., R. A. Geiger, and M. H. Smith. 1973. Accumulation and redistribution
of radiocesium by migratory waterfowl inhabiting a reactor cooling reservoir.
68
Environmental behavior of radionuclides released in the nuclear industry,
International Atomic Energy Agency symposium (IAEA-SM-172/72) Vienna,
Austria, pp. 373-384.
Brisbin Jr., I. L., R. J. Beyers, R. W. Dapson, R. A. Geiger, J. B. Gentry, J. W. Gibbons,
M. H. Smith, and S. K. Woods. 1974. Patterns of radiocesium in the sediments of
a stream channel contaminated by production reactor effluents. Health Physics
27:19-27.
Brisbin, I. L., Jr. and M. H. Smith. 1975. Radiocesium concentrations in whole-body
homogenates and several body compartments of naturally contaminated white-
tailed deer. Mineral Cycling in Southeastern Ecosystems (Howell, F.G., J.B.
Gentry, and M.H. Smith, eds.). ERDA Symposium Series (CONF-740513), pp.
542-556.
Brisbin Jr., I. L., and M. J. Vargo. 1982. Four-year declines in radiocesium
concentrations of American coots inhabiting a nuclear reactor cooling reservoir.
Health Physics 43:266-269.
Brisbin Jr., I. L. 1991. Avian radioecology. Current Ornithology, Volume 8. (Ed. D.M.
Power) Plenum Publishing Corporation, New York, NY, pp. 69-140.
Brisbin Jr., I. L., and R. A. Kennamer. 2000. Long-term studies of radionuclide
contamination of migratory waterfowl at the Savannah River Site: Implications
for habitat management and nuclear waste site remediation. Studies in Avian
Biology 21:57-64.
Bryan Jr., A. L., W. A. Hopkins, J. H. Parikh, B. P. Jackson, and J. M. Unrine. 2012.
Coal fly ash basins as an attractive nuisance to birds: Parental provisioning
69
exposes nestlings to harmful trace elements. Environmental Pollution 161:170-
177.
Carlton, W. H., L. R. Bauer, A. G. Evans, L. A. Geary, C. E. Murphy Jr., J. E. Pinder,
and R. N. Strom. 1992. Cesium in the Savannah River Site environment. No. DE-
AC09-89SR18035 Westinghouse Savannah River Company, Aiken, SC.
Cherry, D. S., S. R. Larrick, R. K. Guthrie, E. M. Davis, and F. F. Sherberger. 1979.
Recovery of invertebrate and vertebrate populations in a coal ash stressed
drainage system. Journal of the Fisheries Research Board of Canada 36:1089-
1096.
Chmielnicka, J., T. Halatek, and U. Jedlinska. 1989. Correlation of cadmium-induced
nephropathy and the metabolism of endogenous copper and zinc in rats.
Ecotoxicology and Environmental Safety 18:268-276.
Christodouleas, J. P., R. D. Forrest, C. G. Ainsley, Z. Tochner, S. M. Hahn, and E.
Glatstein. 2011. Short-term and long-term health risks of nuclear-power-plant
accidents. New England Journal of Medicine, 364:2334-2341.
Clarkson, T. W., L. Magos, and G. J. Meyers. 2003. Human exposure to mercury: The
three modern dilemmas. Journal of Trace Elements in Experimental Medicine
16:321-343.
Coğun, H. Y., T. A. Yüzereroğlu, Ö. Firat, G. Gök, and F. Kargin. 2006. Metal
concentrations in fish species from the Northeast Mediterranean Sea.
Environmental Monitoring and Assessment 121:431-438.
70
Conder, J. M., and J. A. Arblaster. 2016. Development and use of wild game
consumption rates in human health risk assessments. Human and Ecological Risk
Assessment 22:251-264.
Cristol, D. A., L. Savoy, D. C. Evers, C. Perkins, R. Taylor, and C. W. Varian-Ramos.
2012. Mercury in waterfowl from a contaminated river in Virginia. Journal of
Wildlife Management 76:1617-1624.
Currie, L. A. 1968. Limits for qualitative detection and quantitative determination.
Analytical Chemistry 40:586-593.
DeVault, T. L., B. D. Reinhardt, I. L. Brisbin, Jr., and O. E. Rhodes, Jr. 2004. Home
ranges of sympatric black and turkey vultures in South Carolina. Condor 106:706-
711.
Domingo, J. L. 1994. Metal-induced developmental toxicity in mammals: A review.
Journal of Toxicology and Environmental Health 42:123-141.
Dorman, L., J. H. Rodgers Jr., and J. W. Castle. 2010. Characterization of ash-basin
waters from a risk-based perspective. Water, Air, and Soil Pollution 206:175-185.
Dreicer, M., A. Aarkrog, R. Alexakhin, L. Anspaugh, N. P. Arkhipov, and K. J.
Johansson. 1996. Consequences of the Chernobyl accident for the natural and
human environments. One Decade after Chernobyl: Summing up the
Consequences of the Accident. International Atomic Energy Agency (IAEA),
Vienna, Austria, pp. 319-361.
Duchesne, J. F., B. Levesque, D. Gauvin, B. Braune, S. Gingras, and E. Dewailly. 2004.
Estimating the mercury exposure dose in a population of migratory bird hunters in
71
the St. Lawrence River region, Quebec, Canada. Environmental Research 95:207-
217.
Dvořák, P., P. Snášel, and K. Beňová. 2010. Transfer of radiocesium into wild boar meat.
Acta Veterinaria Brno 79:85-91.
Eisler, R. 1998. Copper hazards to fish, wildlife, and invertebrates: A synoptic review.
U.S. Geological Survey, Biological Resources Division, Biological Science
Report USGS/BRD/BSR – 1998-0002, Washington, DC, pp. 120.
Ellis, E. C. and N. Ramankutty. 2008. Putting people in the map: Anthropogenic biomes
of the world. Frontiers in Ecology and the Environment 6:439-447.
European Economic Community (EEC). 1986. Derived reference levels as a basis for the
control of food stuffs following a nuclear accident: A recommendation from the
group of experts set up under Article 31 of the Eurotom Treaty. EEC Regulation
1701/86, Commission of the EEC, Printing Office, Brussels, Belgium.
Evangeliou, N., Y. Balkanski, A. Cozic, and A. P. Møller. 2013. Global Transport and
deposition of 137Cs following the Fukushima nuclear power plant accident in
Japan: Emphasis on Europe and Asia using high–resolution model versions and
radiological impact assessment of the human population and the environment
using interactive tools. Environmental Science and Technology 47:5803-5812.
Evans, D.W., J. J. Alberts, and R. A. Clark. 1983. Reversible ion-exchange fixation of
cesium-137 leading to mobilization from reservoir sediments. Geochimica et
Cosmochimica Acta 47:1041–1049.
Evers, D. C., N. M. Burgess, L. Champoux, B. Hoskins, A. Major, W. M. Goodale, R. J.
Taylor, R. Poppenga, T. Daigle. 2005. Patterns and interpretation of mercury
72
exposure in freshwater avian communities in northeastern North America.
Ecotoxicology 14:193-221.
Farag, A. M., D. D. Harper, L. Cleveland, W. G. Brumbaugh, and E. E. Little. 2006. The
potential for chromium to affect the fertilization process of chinook salmon
(Oncorhynchus tshawytscha) in the Hanford Reach of the Columbia River,
Washington, USA. Archives of Environmental Contamination and Toxicology
50:575-579.
Farkas, A., J. Salánki, and A. Specziár. 2003. Age- and size-specific patterns of heavy
metals in the organs of freshwater fish Abramis brama L. populating low-
contaminated site. Water Research 37:959-964.
Fendley, T. T., M. N. Manlove, and I. L. Brisbin Jr. 1977. The accumulation and
elimination of radiocesium by naturally contaminated wood ducks. Health Physics
32: 415-422.
Gaines, K. F., C. S. Romanek, C. S. Boring, C. G. Lord, M. Gochfeld, and J. Burger.
2002. Using raccoons as an indicator species for metal accumulation across
trophic levels: A stable isotope approach. Journal of Wildlife Management.
66:811-821.
Gilbert, R. O., and R. R. Kinnison. 1981. Statistical methods for estimating the mean and
variance from radionuclide data sets containing negative, unreported or less-than
values. Health Physics 40:377-390.
Goyer, R. A. 1991. Toxic effects of metals. Casarett and Doull’s Toxicology, Fourth
edition. (Eds. C.D. Klassen, M.O. Amdur, J. Doull) Macmillan, New York, NY,
pp. 623-680.
73
Grandjean, P., R. F. White, A. Nielsen, D. Cleary, and E. C. Santos. 1999.
Methylmercury neurotoxicity in Amazonian children downstream from gold
mining. Environmental Health Perspectives 107:587-591.
Gudiksen, P. H., T. F. Harvey, and R. Lange. 1989. Chernobyl source term, atmospheric
dispersion, and dose estimation. Health Physics 57:697-706.
Hall, B. D., L. A. Baron, and C. M. Somers. 2009. Mercury concentrations in surface
water and harvested waterfowl from the prairie pothole region of Saskatchewan.
Environmental Science and Technology, 43:8759-8766.
Havelková, M., L. Dušek, D. Némethová, G. Poleszczuk, and Z. Svobodová 2008.
Comparison of mercury between liver and muscle−A biomonitoring of fish from
lightly and heavily contaminated localities. Sensors 8:4095-4109.
Heinz, G. H. 1996. Selenium in birds. Environmental Contaminants in Wildlife:
Interpreting Tissue Concentrations. (Eds. W. N. Beyer, G.H. Heinz, and A. W.
Redmon-Norwood) CRC Press, Boca Raton, FL, pp. 447-458.
Heinz, G. H., and D. J. Hoffman. 1998. Methylmercury chloride and selenomethionine
interactions on health and reproduction in mallards. Environmental Toxicology
and Chemistry 17:139-145.
Hinton, T. G. 1998. Estimating human and ecological risks from exposure to radiation.
Risk Assessment: Logic and Measurement (Eds. M. C. Newman, and C. L.
Strojan) Ann Arbor Press, Chelsea, MI, pp. 143-166.
Hinton, T. G., R. Alexakhin, M. Balonov, N. Gentner, J. Hendry, B. Prister, P. Strand,
and D. Woodhead. 2007. Radiation-induced effects on plants and animals:
Findings of the United Nations Chernobyl Forum. Health Physics 93:427-440.
74
Holben, D. H. 2002. Selenium content in vension, squirrel, and beef purchased or
produced in Ohio, a low selenium region of the United States. Journal of Food
Science 67:431-433.
Hopkins, W. A., M. T. Mendonca, C. L. Rowe, J. D. Congdon. 1998. Elevated trace
element concentrations in southern toads Bufo terrestris, exposed to coal
combustion wastes. Archives of Environmental Contamination and Toxicology
35:325-329.
Hopkins, W. A., C. L. Rowe, and J. D. Congdon. 1999. Elevated trace element
concentrations and standard metabolic rate in banded water snakes (Nerodia
fasciata) exposed to coal combustion wastes. Environmental Toxicology and
Chemistry 18:1258-1263.
Hopkins, W. A., J. D. Congdon, and J. K. Ray. 2000. Incidence and impact of axial
malformations in larval bullfrogs (Rana catesbeiana) developing in sites polluted
by a coal-burning power plant. Environmental Toxicology and Chemistry 19:862-
868.
Horton, J. H. 1974. Technical Division, Savannah River Laboratory. Memorandum to
B.C. Rusche. Subject: Mercury in the separations areas seepage basins. DPST-74-
231. SRS Phase II Database GKH.
Humphries, M. 1999. Global climate change: Coal use in China and other Asian
developing countries. Washington, D.C. Congressional Research Service, the
Library of Congress. University of North Texas Digital Library
http://digital.library.unt.edu/ark:/67531/metacrs848/. Accessed 11 January 2016.
75
Hughes, M. F., B. D. Beck, Y. Chen, A. S. Lewis, and D. J. Thomas. 2011. Arsenic
exposure and toxicology: A historical perspective. Toxicological Sciences
123:305-332.
Ikem, A., N. O. Egiebor, and K. Nyavor. 2003. Trace elements in water, fish, and
sediment from Tuskegee Lake Southeastern USA. Water, Air, and Soil Pollution
149:51-75.
Kennamer, R. A., I. L. Brisbin Jr., C. D. McCreedy, and J. Burger. 1998. Radiocesium in
Mourning Doves: Effects of a contaminated reservoir drawdown and risk to
human consumers. Journal of Wildlife Management. 62:497-508.
Kennamer, R. A. 2003. Recoveries of ring-necked ducks banded on the U.S. Department
of Energy’s Savannah River Site, South Carolina. The Oriole 68:8-14.
Lemly, A. D. 1996. Selenium in aquatic organisms. Environmental Contaminants in
Wildlife: Interpreting Tissue Concentrations. (Eds. W. N. Beyer, G. H. Heinz, and
A. W. Redmon-Norwood) CRC Press, Boca Raton, FL, pp. 427-455.
Lemly, A. D. 2002. Symptoms and implications of selenium toxicity in fish: the Belews
Lake case example. Aquatic Toxicity 57:39-49.
Lemly, A. D., and J. Skorupa. 2012. Wildlife and coal waste policy debate: Proposed
rules for coal waste disposal ignore lessons from 45 years of wildlife poisoning.
Environmental Science and Technology 46:8595-8600.
Lide, R. F. 1994. Carolina bays and similar natural depression wetlands of the SRS. GIS
coverage in ARC/INFO. Savannah River Ecology Laboratory, Savannah River
Site, Aiken, SC.
76
Luther, L. 2010a. Managing coal combustion waste (CCW): Issues with disposal and use.
Congressional Research Service. CRS report for Congress, 7-5700, R40544, pp.
1-26. https://www.fas.org/sgp/crs/misc/R40544.pdf. Accessed 29 January 2016.
Luther, L. 2010b. Regulating Coal Combustion Waste Disposal: Issues for Congress.
Congressional Research Service. CRS report for Congress, 7-5700, R41341, pp.
1-22. http://www.fas.org/sgp/crs/misc/R41341.pdf. Accessed 29 January 2016.
Mason, R. P., J-M Laporte, and S. Andres. 2000. Factors controlling the bioaccumulation
of mercury, methylmercury, arsenic, selenium, and cadmium by freshwater
invertebrates and fish. Archives of Environmental Contamination and Toxicology
38:283-297.
Masson, O., A. Baeza, J. Bieringer, K. Brudecki, S. Bucci, M. Cappai, F. P. Carvalho et
al. 2011. Tracking of airborne radionuclides from the damaged Fukushima Dai-
ichi nuclear reactors by European networks. Environmental Science and
Technology 45:7670-7677.
Masson, O., W. Ringer, H. Malá, P. Rulik, M. Dlugosz-Lisiecka, K. Eleftheriadis, O.
Meisenberg, A. De Vismes-Ott, and F. Gensdarmes. 2013. Size distributions of
airborne radionuclides from the Fukushima nuclear accident at several places in
Europe. Environmental Science and Technology 47:10995-11003.
Mayer, J. J., R. A. Kennamer, and R. T. Hoppe. 1986. Waterfowl of the Savannah River
Plant. Final Report. Report SREL-22 UC-66e. Savannah River Ecology
Laboratory, Division of Stress and Wildlife Ecology, Aiken, SC, USA.
Mayfield, D. B., S. Thakali, W. T. Mehler, A. S. Lewis. 2013. Ecological effects of coal
combustion products (CCPs). A literature review of observed effects and
77
considerations for managing risks. 2013 World of Coal Ash Conference,
Lexington, Kentucky. http://www.flyash.info/. Accessed 22 February 2016.
McCloskey, J. T., and M. C. Newman. 1995. Sediment preference in the Asiatic clam
(Corbicula fluminea) and viviparid snail (Campeloma decisum) as a response to
low-level metal and metalloid contamination. Archives of Environmental
Contamination and Toxicology 28:195-202.
Mishra, U. 1990. Comparison of radionuclides levels from the Chernobyl reactor accident
and from global fallout. Journal of Radioanalytical and Nuclear Chemistry
138:119-125.
Mohler, H. J., F. W. Whicker, and T. G. Hinton. 1997. Temporal trends of 137Cs in an
abandoned reactor cooling reservoir. Journal of Environmental Radioactivity
37:251-268.
Neely, W. B. 1980. Chemicals in the environment: Distribution, transport, fate, analysis.
Marcel Dekker, NY, pp. 245.
Newman, M. C., P. M. Dixon, B. B. Looney, and J. E. Pinder, III. 1989. Estimating mean
and variance for environmental samples with below detection limit observations.
Water Resources Bulletin 25:905-916.
Ohlendorf, H. M., and W. J Fleming. 1988. Birds and environmental contaminants in San
Francisco and Chesapeake Bays. Marine Pollution Bulletin 19:487-495.
Ohlendorf, H. M. and G. H. Heinz. 2011. Selenium in birds. Environmental
Contaminants in Biota: Interpreting Tissue Concentrations (Eds. N. Beyer and
J.M. Meador), CRC Press, Boca Raton, FL, pp. 669-701.
78
Omojola, A. B. 2007. Carcass and organoleptic characteristics of duck meat as influenced
by breed and sex. International Journal of Poultry Science 6:329-334.
Otter, R. R., F. C. Bailey, A. M. Fortner, and S. M. Adams. 2012. Trophic status and
metal bioaccumulation differences in multiple fish species exposed to coal ash-
associated metals. Ecotoxicology and Environmental Safety 85:30-36.
Paller, M. H., J. W. Littrell, and E. L. Peters. 1999. Ecological half-lives of 137Cs in fishes
from the Savannah River Site. Health Physics 77:392-407.
Peles, J. D., A. L. Bryan Jr., C. T. Garten Jr., D. O. Ribble, and M. H. Smith. 2000.
Ecological half-life of 137Cs in fish from a stream contaminated by nuclear reactor
effluents. Science of the Total Environment 263:255-262.
Penglase, S., K. Hamre, and S. Elligen. 2014. Selenium and mercury have a synergistic
negative effect on fish reproduction. Aquatic Toxicology Journal 149:16-24.
Peterson, S. A., N. V. Ralston, D. V. Peck, J. V. Sickle, J. D. Robertson, V. L. Spate, and
J. S. Morris. 2009. How might selenium moderate the toxic effects of mercury in
stream fish of the western US? Environmental Science and Technology 43:3919-
3925.
Potter, C. M., Brisbin, I. L., Jr., McDowell, S. G., Whicker, F. W., 1989. Distribution of
137Cs in the American coot (Fulica americana). J. Environmental Radioactivity 9,
105-115.
Reash, R. J. 2012. Selenium, arsenic, and mercury in fish inhabiting a fly ash exposure
gradient: Interspecific bioaccumulation patterns and elemental associations.
Environmental Toxicology and Chemistry 31:739-747.
79
Rice, K. M., E. M. Walker Jr., M. Wu, C. Gillette, and E. R. Blough. 2014.
Environmental mercury and its toxic effects. Journal of Preventative Medicine
and Public Health 47:74-83.
Rigg, D. K., M. N. Wacksman, J. Iannuzzi, T. F. Baker, M. Adams, and M. S. Greeley Jr.
2015. Assessing ecological risks to the fish community from residual coal fly ash
in Watts Bar Reservoir, Tennessee. Integrated Environmental Assessment and
Management 11:88-101.
Rowe, C. L., O. M. Kinney, A. P. Fiori, and J. D. Congdon. 1996. Oral deformities in
tadpoles (Rana catsbeiana) associated with coal ash deposition: Effects on
grazing ability and growth. Freshwater Biology 36:723-730.
Rowe, C. L., W. A. Hopkins, J. D. Congdon. 2002. Ecotoxicological implications of
aquatic disposal of coal combustion residues in the United States: A review.
Environmental Monitoring and Assessment 80:207-276.
Ruhl, L., A. Vengosh, G. S. Dwyer, H. Hsu-Kim, A. Deonarine, M. Bergin, and J.
Kravchenko. 2009. Survey of environmental and health impacts in the immediate
aftermath of the coal ash spill in Kingston, Tennessee. Environmental Science and
Technology 43:6326-6333.
Ruhl, L., A. Vengosh, G. S. Dwyer, H. Hsu-Kim, G. Schwartz, A. Romanski, and S. D.
Smith. 2012. The impact of coal combustion residue effluent on water resources:
A North Carolina example. Environmental Science and Technology 46:12226-
12233.
80
Sanderson, E. W., M. Jaiteh, M. A. Levy, K. H. Redford, A. V. Wannebo, and G.
Woolmer. 2002. The human footprint and the last of the wild. BioScience,
52:891-904.
Savannah River Nuclear Solutions, LLC. 2011. Environmental, safety, health, and quality
regulatory integration and environmental services. Environmental Management
System (EMS) description manual.
http://www.srs.gov/general/pubs/envbul/documents/ems_manual.pdf Accessed
15 October 2014.
Scheuhammer, A. M. 1987. The chronic toxicity of aluminum, cadmium, mercury and
lead in birds. Environmental Pollution 46:263-295.
Shore, R. F., M. G. Pereira, L. A. Walker and D. R Thompson. 2011. Mercury in non-
marine birds and mammals. Environmental Contaminants in Biota: Interpreting
Tissue Concentrations (Eds. N. Beyer & J. M. Meador), CRC Press, Boca Raton,
FL, pp. 609-624.
Skuterud, L., E. Gaare, I. M. Eikelmann, K. Hove, and E. Steinnes. 2005. Chernobyl
radioactivity persists in reindeer. Journal of Environmental Radioactivity 83:231-
252.
Stepanauskas, R., T. C. Glenn, C. H. Jagoe, R. C. Tuckfield, A. H. Lindell, and J. V.
McArthur. 2005. Elevated microbial tolerance to metals and antibiotics in metal-
contaminated industrial environments. Environmental Science and Technology
39:3671-3678.
81
Taggart, M. A., R. Mateo, J. M. Charnock, F. Bahrami, A. J. Green, and A. A. Meharg.
2009. Arsenic rich iron plaque on macrophyte roots – An ecological risk?
Environmental Pollution 157:946-954.
Taggart, M. A., M. M. Reglero, P. R. Camarero, R. Mateo. 2011. Should legislation
regarding maximum Pb and Cd levels in human food also cover large game meat?
Environmental International 37:18-25.
United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR).
1988. Sources, effects, and risks of ionizing radiation. United Nations, NY, pp.
647.
United States Department of Energy (USDOE). 2012.
http://energy.gov/em/articles/savannah-river-site-retires-coal-fired-d-area-
powerhouse-after-nearly. Accessed 8 September 2015.
United States Environmental Protection Agency (USEPA). 1998. Integrated risk
information system (IRIS) on arsenic. National Center for Environmental
Assessment, Office of Research and Development, Washington, DC.
http://www.epa.gov/ttnatw01/hlthef/arsenic.html#ref5. Accessed 30 January
2015.
United States Environmental Protection Agency (USEPA). 1999a. Integrated risk
information system (IRIS) on methylmercury. National Center for Environmental
Assessment, Office of Research and Development, Washington, DC.
http://www.epa.gov/ttnatw01/hlthef/mercury.html#ref13. Accessed 30 January
2015.
82
United States Environmental Protection Agency (USEPA). 1999b. Integrated risk
information system (IRIS) on selenium and compounds. National Center for
Environmental Assessment, Office of Research and Development, Washington,
DC. http://www.epa.gov/airtoxics/hlthef/selenium.html#ref4. Accessed 30
January 2015.
United States Environmental Protection Agency (USEPA). 2001. Water quality criterion
for the protection of human health: Methylmercury. Office of Science and
Technology, Office of Water, Washington DC.
http://water.epa.gov/scitech/swguidance/standards/criteria/health/upload/2009_01
_15_criteria_methylmercury_mercury-criterion.pdf Accessed 15 January 2015.
United States Environmental Protection Agency (USEPA). 2012a. Mercury and air toxics
standards, cleaner power plants. U.S. Environmental Protection Agency,
Washington, DC. http://www.epa.gov/mats/powerplants.html. Accessed 07
January 2016.
United States Environmental Protection Agency (USEPA). 2012b. Waste and cleanup
risk assessment glossary. U.S. Environmental Protection Agency Office of Solid
Waste Emergency Response, Washington, DC.
http://www.epa.gov/oswer/riskassessment/glossary.htm. Accessed 16 October
2014.
United States Environmental Protection Agency (USEPA). 2014. Mercury health effects.
Washington, DC. http://www.epa.gov/mercury/effects.htm. Accessed 3
November 2014.
United States Environmental Protection Agency (USEPA). 2015. Frequent questions
83
about coal ash disposal rule. U.S. Environmental Protection Agency, Washington,
DC. http://www.epa.gov/coalash/frequent-questions-about-coal-ash-disposal-
rule#3. Accessed 07 January 2016.
United States Fish and Wildlife Service (USFWS) and United States Census Bureau
(USCB). 2011. 2011 National survey of fishing, hunting and wildlife associated
recreation. U.S. Department of the Interior, Washington, DC.
https://www.census.gov/prod/2012pubs/fhw11-nat.pdf. Accessed 09 February
2015.
Unrine, J. M., W. A. Hopkins, C. S. Romanek, and B. P. Jackson. 2007.
Bioaccumulation of trace elements in omnivorous amphibian larvae: Implications
for amphibian health and contaminant transport. Environmental Pollution
149:182-192.
Van Dyke, J. U., W. A. Hopkins, and B. P. Jackson. 2013. Influence of relative trophic
position and carbon source on selenium bioaccumulation in turtles from a coal
fly-ash spill site. Environmental Pollution 182:45-52.
Vitousek, P. M., H. A. Mooney, J. Lubchenco, and J. M. Melillo. 1997. Human
domination of Earth’s ecosystems. Science 277:494-499.
Wiener, J. G., D. P. Krabbenhoft, G. H. Heinz, A. M. Scheuhammer. 2003.
Ecotoxicology of mercury. Handbook of Ecotoxicology, Second edition. (Eds. D.
J. Hoffman, B. A. Rattner, G.A. Burton Jr., J. Cairns Jr.) CRC Press, Boca Raton,
FL, pp 409-463.
White, D. L., and K. E. Gaines. 2000. The Savannah River Site: Site description, land
use, and management history. Studies in Avian Biology 21:8-17.
84
Wike, L. D., F. D. Martin, E. A. Nelson, N. V. Halverson, J. J. Mayer, M. H. Paller, R. S.
Riley, M. G. Serrato, and W. L. Specht, 2006. SRS Ecology Environmental
Information Document. Westinghouse Savannah River Company. WSRC-TR-
2005-00201.
Wolkers, H., T. Wensing, and G. W. Bruinderink. 1994. Heavy metal contamination in
organs of red deer (Cervus elaphus) and wild boar (Sus scrofa) and the effect on
some trace elements. Science of the Total Environment 144:191-199.
Wright, M. S., G. L. Peltier, R. Stepanauskas, and J. V. McArthur. 2006. Bacterial
tolerance to metals and anti-biotic in metal-contaminated and reference streams.
FEMS Microbiology Ecology 58:293-302.
Yang, D.-Y., Y.-W., J. M. Gunn, and N. Betzille. 2010. Inverse relationships between
selenium and mercury in tissues of young walleye (Stizosedion vitreum) from
Canadian boreal lakes. Science of the Total Environment 408:1676-1683.
Yudovich, Y. E., and M. P. Ketris. 2005a. Arsenic in coal: A review. International
Journal of Coal Geology 61:141-196.
Yudovich, Y. E., and M. P. Ketris. 2005b. Mercury in coal: A review. Part 1.
Geochemistry. International Journal of Coal Geology 62:107-134.
Yudovich, Y. E., and M. P. Ketris. 2006. Selenium in coal: A review. International
Journal of Coal Geology 67:112-126.
85
CHAPTER 3
WATERFOWL EXPOSURE TO COAL COMBUSTION WASTES AND HUMAN
CONSUMPTION RISKS 1
____________________
1 Oldenkamp, R. E., A. L. Bryan, Jr., R. A. Kennamer, and J. C. Beasley. To be submitted to the Journal of Wildlife Management.
86
ABSTRACT
Waterfowl are important and popular game species and thus represent a potential
pathway for human exposure to anthropogenic pollutants through the consumption of
contaminated meat. In particular, settling basins containing coal combustion waste
(CCW) enriched in trace elements such as arsenic (As), selenium (Se), and mercury (Hg)
are often utilized by free-ranging migratory waterfowl, representing potential sources for
contaminant uptake. We developed an experiment to restrict waterfowl to a CCW
contaminated basin then quantified levels of contaminants in waterfowl and modeled
trace element burdens in blood, muscle, and liver for known time periods of exposure
(between 3 and 92 days of exposure). We developed equations to predict muscle/liver
burdens based on concentrations in blood as a potential non-destructive sampling method
and used muscle tissue to calculate human consumption limits based on concentrations of
recognized elements of human health concern (As, Se, and Hg). We observed a
significant increase in Se concentrations in muscle, liver, and blood tissues over the
duration of our experiment. Consumption limits of waterfowl breast muscle based on
EPA chronic oral reference dose guidelines for As, Se, and Hg. Consumption limits were
lowest for As, with an allowance as low as 2.9 meals per month, although given the lack
of As accumulation observed through time these results may reflect elevated levels of As
for ring-necked ducks prior to the initiation of this study. Consumption limits based on
observed Se accumulation decreased from 26.7 to 6.2 meals per month during the course
of this study. Children’s allowances based on average concentrations were approximately
half that of adults for each element. These data provide unique insights into accumulation
rates of contaminants for waterfowl utilizing habitats contaminated with CCW and
87
suggest more comprehensive, long-term monitoring of contaminant burdens in waterfowl
is needed across a broad geographic scale to assess waterfowl exposure (and human
hunter exposure risk) to pollutants.
INTRODUCTION
Waterfowl are an important game species globally. In the U.S. alone ~2.5 million
hunters spend nearly $1.8 billion on waterfowl hunting annually (USFWS and USCB
2011). The vast majority of hunters pursue game for consumption, and waterfowl hunters
consume numerous birds each year due to large daily federal bag limits (daily max
allowed during hunting season; Duchesne et al. 2004; Smith et al. unpublished
manuscript). Despite the widespread consumption of waterfowl and other game, wild
game is not subjected to the same regulatory testing as livestock and thus consumption of
free-ranging wildlife could potentially expose unknowing hunters, their families, and
recipients of donated game to environmental pollution (Cristol et al. 2012, Conder and
Arblaster 2016). Waterfowl, in particular, could represent an important pathway for
contaminant exposure in humans because they are vulnerable to uptake of contaminants
due to extensive use of aquatic habitats, where many pollutants occur, and foraging
behaviors that disturb sediments (Bryan et al. 2012). Moreover, waterfowl are highly
mobile and migratory and thus can transport pollutants hundreds or thousands of
kilometers from point sources (Kennamer 2003, Cristol et al. 2012, Conder and Arblaster
2016).
Although numerous pollutants exist within aquatic ecosystems, surface
impoundments containing coal combustion wastes (CCW) represent a potential pathway
for wildlife exposure to several trace elements flagged by the EPA as environmental and
88
human health risks (Luther 2010a,b; Rowe et al. 2002). Millions of tons of CCW are
deposited into surface impoundment ponds across the US and as of 2015 the EPA
identified more than 500 coal power facilities with 735 surface impoundments to store
CCW in the US alone (USEPA 2015). Furthermore, coal use is on the rise in developing
countries, making CCW important at a global scale (Humphries 1999). Many of these
storage areas are unlined and present a risk of environmental pollution from the
multifarious combustion byproducts that are enriched in potentially toxic trace elements
such as arsenic (As), selenium (Se), and mercury (Hg; Greeley et al. 2016). Indeed, more
than 20 surface impoundment accidents have caused substantive fish, wildlife, and
economic losses (Lemly 1996, 2002; Ruhl et al. 2009; Lemly and Skorupa 2012; Rigg et
al. 2015). This does include exposure at these impoundments when there is not a
structural failure or release of effluents into nearby waterways. Waterfowl and waterbirds
may be particularly susceptible to CCW exposure as open surface impoundments often
have wooded or herbaceous areas of cover along edges supporting a diversity of prey
(e.g., aquatic vegetation, insects, amphibians, and fish; Rowe et al. 2002). Yet,
comparatively little is known about exposure of waterfowl to CCW compared to fish,
amphibians, and reptiles (Hopkins et al. 1999, 2000; Yang et al. 2010; Van Dyke 2013).
Bioaccumulation, the build-up of contaminants within the tissues of the body, can
occur through the consumption of the source pollutant or through movement in trophic
levels within the food web (USEPA 2012b). Trace elements in or around CCW storage
areas have been found to bioaccumulate or biomagnify in wildlife, especially aquatic or
semi-aquatic organisms (Yudovich and Ketris 2005a,b; Yudovich and Ketris 2006; Reash
2012; Otter et al. 2012), posing potential health risks to aquatic organisms, terrestrial
89
wildlife, and humans (Dorman et al. 2010, Lemly and Skorupa 2012, Ruhl et al. 2012,
Mayfield et al. 2013, Rice et al. 2014). Examples of CCW pollution affecting wildlife are
well documented. For instance, coal fly ash effluent discharged into Belews Lake and
Hyco Reservoir in North Carolina and in Martin Creek Reservoir in Texas have resulted
in trace element accumulation in aquatic biota, with Se contamination resulting in
significant increases in developmental abnormalities in fish larvae, salamanders, and toad
tadpoles and local extinctions of sensitive species (Lemly 1996, 2002; Hopkins et al.
2006; Roe et al. 2006; Metts et al. 2013). The largest industrial waste spill in U.S. history,
from Kingston (TVA) plant in Tennessee, resulted in elevated levels of As, Se, and Hg in
fish and resident raccoons, signifying transfer of contamination to terrestrial organisms
(Ruhl et al. 2009, Beck et al. 2013, Van Dyke et al. 2013, Beck et al. 2015, Rigg et al.
2015).
Most laboratory toxicology studies that have sought to identify lethal doses of
contaminants and measure morbidity, used liver tissue as the standard to explain health
effects of exposure. Liver and kidney, accumulate substantially greater concentrations of
most trace elements than muscle tissue (Farkas et al. 2003, Ikem et al. 2003, Coğun et al.
2006). While exceptions exist, (Mason et al. 2000, Havelková et al. 2008), this has meant
that detoxifying organs have been used most often as biological indicators of contaminant
presence in water. However, this common approach in laboratory toxicology studies
typically ignores element concentrations in muscle tissue and thus is insufficient to
facilitate calculations of human exposure risk for consumption of wildlife inhabiting
contaminated ecosystems. This approach also ignores complications with natural
exposure (e.g. fluctuations in exposure through time and space and through food items
90
with differing levels of contamination) and multiple contaminant exposure scenarios (i.e.
coal ash has many different trace elements of differing toxicity that may interact) that are
the norm in polluted ecosystems.
Data on contaminant concentrations in wild waterfowl is surprisingly scarce
despite their importance as a consumed game species and their potential to accumulate
trace elements or other contaminants through feeding in sediments of aquatic habitats
where pollutants often concentrate (Chapter 2, see Cristol et al. 2012). Moreover,
interpretations of contaminant levels in many studies can be problematic because wild
birds can potentially transition between contaminated and uncontaminated habitats during
both migratory and resident periods. Testing of wild birds from contaminated areas can
provide a snapshot of contaminant levels within tissues; however, the length of time birds
were present within the contaminated habitat is generally unknown and thus experimental
research is needed to elucidate accumulation rates of contaminants to inform observed
concentrations in free-ranging waterfowl (Chapter 2). In this study we characterized
accumulation rates of several trace elements of importance to human and wildlife health
in waterfowl restricted to a coal fly ash surface impoundment over a 3-month period, and
use these data to assess potential risks to waterfowl and human hunters consuming
waterfowl that have utilized CCW impoundment facilities for varying periods of time.
Specifically, our objectives were to 1) quantify trace element uptake in blood, muscle,
and liver tissues over known periods of time by waterfowl exposed in situ to a coal ash
settling basin and elucidate potential interactions among elements, 2) develop a model to
predict muscle/liver burdens based on concentrations in blood as a potential non-
destructive sampling method and test the performance of the model against a subset of
91
our data, and 3) calculate human consumption limits based on concentrations of
recognized elements of human health concern (As, Se, and Hg) over known time periods
of exposure.
METHODS
Study Area
This study occurred on the Savannah River Site (SRS), a ~800 km2 limited access
former nuclear production and research facility owned and operated by the U.S.
Department of Energy. The SRS is located in the coastal plain of South Carolina and is
classified as a National Environmental Research Park (White and Gaines 2000). Created
in 1951 to provide nuclear weapons materials at the beginning of the Cold War
(Savannah River Nuclear Solutions, LLC 2011), the SRS now contains five
decommissioned nuclear reactors, radioactive materials processing facilities, and nine
retired coal power plants (White and Gaines 2000). These facilities encompass <5% of
the SRS; the remaining area is comprised of a mosaic of natural habitats consisting of
managed pine stands (54%), wetlands (23%), upland hardwood and mixed forest (11%),
grasslands (9%), and upland scrub forest (3%; Lide 1994, White and Gaines 2000,
DeVault et al. 2004).
The ash basins associated with the SRS’ D-area coal fired power plant represent
the most thoroughly studied ash disposal system in the world. From 1953-2012 the D-
Area coal-fired power plants on the SRS were operational and during this time sluiced fly
ash was deposited into settling impoundments that flowed into Beaver Dam Creek, a
tributary of the Savannah River, and nearby wetlands (Halverson et al. 1997, Gaines et al.
2002, USDOE 2012). These settling impoundments are unlined earthen basins located
92
approximately 0.5 km from the Savannah River. Vegetation has grown in and around
edges of the basins and several species of waterfowl are frequently observed foraging in
these artificial habitats (Chapter 2). In particular, following cessation of discharge, basin
1 (15 ha), the focal area of this research, became partially filled and vegetation expanded
inward from already established wooded edges creating a small semi-vegetated wetland
enclosed by the rest of the basin (Figure 3-1). Numerous studies have reported
bioaccumulation and adverse effects from exposure to elevated levels of aluminum (Al),
As, cadmium (Cd), chromium (Cr), iron (Fe), Hg, manganese (Mn), nickel (Ni), Se, and
zinc (Zn) for a wide array of organisms inhabiting the basins and creek watershed
including bacteria, aquatic invertebrates, amphibians, fish, turtles, alligators, and birds
(Cherry et al. 1979; Hopkins et al. 1999, 2000; Rowe et al. 1996, 2002; Stepanauskas et
al. 2005; Hopkins et al. 2006; Roe et al. 2006; Wright et al. 2006; Bryan et al. 2012;
Metts et a;. 2013). There are abundant populations of waterfowl that overwinter or are
resident on SRS water bodies and several species are commonly observed using the D-
area CCW impoundments (Chapter 2, Mayer et al. 1986, Kennamer 2005).
Trapping and Sample Collection
Male ring-necked ducks were selected as a target species for this research because
of their extensive use of the SRS as a stop-over and over-wintering location during
migration (Mayer et al. 1986) and because of their varied diet, which includes aquatic
vegetation, insects, snails, and mussels (Hoppe et al. 1986). In December 2014-February
2015 ducks were trapped with swim-in box traps baited with corn placed at L-lake, an
uncontaminated waterbody on the SRS. Upon capture, ducks were banded with uniquely
numbered USFWS bands and fitted with a colored nasal saddle to distinguish unique
93
cohorts of birds during collections. Approximately 1 ml of blood was collected from the
basilica or brachial vein in the wing, or less often the metatarsal vein in the leg, and
placed into tubes without anticoagulant and allowed to clot before being frozen at -80°C
for later trace element analysis. This blood served as a baseline level of contamination to
determine if any birds had high levels of any trace elements in their blood prior to our
experiment, indicating possibility of accumulation in the muscle and liver from previous
locations along their migratory routes. All birds were translocated to ash basin 1 (Figure
3-1) where we clipped the ends of all flight feathers on one wing prior to release to
prevent them from leaving the basin but allowing them to freely move around the basin
and forage.
After release at the ash basin, ducks were lethally collected (with a shotgun)
periodically, aiming to spread out the collections every few days between 3 and 92 days
of exposure. After collection, we collected a 1 ml blood sample via cardiac puncture and
placed samples into tubes without anticoagulant that we subsequently stored in a -80°C
freezer for later trace element analysis. We also collected a weight for each bird and froze
them at -20°C for later dissection. All animal handling practices and euthanasia were
carried out with accompanying federal and state collecting permits and in accordance
with University of Georgia Animal Care and Use guidelines under protocol A2013 06-
004-Y3-A1. After thawing, we dissected all ducks to collect breast muscle and liver
tissues for trace element analyses. We collected wet weights for these tissue samples,
which we subsequently freeze-dried and re-weighed prior to homogenizing them into a
powder using a coffee grinder. We cleaned grinder canisters with a 5% nitric acid
solution and dried canisters between uses.
94
Elemental Analysis
We conducted analyses for trace elements [V, Cr, Ni, Cu, Zn, As, Se, Cd, and Pb]
and total mercury (THg) content on muscle, liver, and blood samples. For trace element
analysis approximately 250 mg of dry muscle or liver sample was microwave digested
(MARSX Xpress, CEM Corporation, Matthews, NC) with 10.0 ml trace metal-grade
nitric acid (70% HNO3). Following digestion, samples were brought to a final volume of
15.0 ml with Milli-Q (18MΩ) water before spectroscopy (Nexlon 300X ICP-MS; Perkin
Elmer, Norwalk, CT) according to QA/QC protocols outlined in EPA Method 6020A
(USEPA 2007). The minimum detection limits (ppm) for each element for muscle and
liver was: V (0.03), Cr (0.03), Ni (0.04), Cu (0.11), Zn (0.11), As (0.05), Se (0.33), Cd
(0.04), and Pb (0.04).
For trace element analysis of blood, we placed approximately 0.50 ml of blood
into a trace metal free tube and weighed samples before placing them in an -80°C freezer
for a few hours to make sure the sample was solid before transferring to the freeze-drier.
After freeze-drying for 12 hours the sample was weighed again before being placed into a
sand bath (~75-85°C) with 2 ml of trace metal-grade nitric acid (70% HNO3) and 1 ml of
hydrogen peroxide (30% H202) added slowly over 90 minutes while still on the heat.
Samples were brought to a final volume of 5.0 ml with Milli-Q (18MΩ) water before
spectroscopy with the same protocol as tissue samples. For blood samples the minimum
detection limits (ppm) for each element was: V (0.06), Cr (0.09), Ni (0.04), Cu (0.25), Zn
(2.15), As (0.05), Se (0.38), Cd (0.04), and Pb (0.47). For quality control purposes,
certified reference material (TORT-3 lobster hepatopancreas; National Research Council,
Ottawa, ON, Canada), a blank, and a digestion replicate were run for every 20 samples.
95
Mean percent recoveries ranged from 86-222% for tissues and 77-146% for blood for
elements in certified reference materials and all element concentrations are presented as
parts per million (ppm) on a dry mass basis.
Following U.S. Environmental Protection Agency (EPA) method 7473 to analyze
total mercury (THg), 30-50 mg subsamples of the freeze-dried/homogenized tissues, and
50 uL of digested, diluted blood, were analyzed by thermal decomposition, catalytic
conversion, amalgamation, and atomic absorption spectrophotometry (DMA 80;
Milestone, Shelton, CT, USA). The instrument detection limit (IDL) for this method was
0.01 nanograms (ng) of total mercury. Quartz sample boats were utilized for the acid-
digested blood samples. For each set of 10 samples we included a replicate, blank, and
two standard reference materials (SRMs; TORT-3 lobster hepatopancreas, and PACS-2
marine sediment, National Research Council of Canada, Ottawa, ON) to ensure quality
assurance and solid SRMs were used to calibrate the instrument. For muscle and liver
tissues method detection limits (MDLs; threefold the standard deviation of procedural
blanks) averaged 0.0004 ppm dry mass; mean percent recoveries of THg for the SRMs
TORT-3 and PACS-2 were 93.6 ± 5.3 and 101.6 ± 3.9 respectively. Blood MDLs
averaged 0.00004 ppm dry mass with mean percent recovery of THg in the SRM, TORT-
3, of 91.2 ± 15.1. Concentrations for tissues and blood are presented as parts per million
(ppm) on a dry mass basis.
Statistical Analysis
Any ducks that exhibited pre-exposure As, Se, or Hg levels two standard
deviations above the mean were excluded from all analyses as they were assumed to
potentially have elevated burdens of these elements prior to translocation to the ash basin.
96
Trace element concentrations have been shown to differ between tissue types in previous
toxicology studies; therefore, we performed tests on muscle and liver samples separately
(Scheuhammer 1987, Boening 2000, Mason et al. 2000, Farkas et al. 2003, Ikem et al.
2003, Coğun et al. 2006, Havelková et al. 2008). For each tissue, when >50% samples
were below detection limits (BDL) for individual elements those elements were excluded
from subsequent analyses. When <50% of samples of a particular tissue were BDL, we
replaced them with 50% of the respective minimum detection limit (MDL; Hall et al.
2009, Fletcher et al. 2014). We tested distributions of all elements for normality (Shapiro-
Wilk test p<0.05) in R (R Core Team 2012) and subsequently transformed data prior to
inclusion in analyses; specific transformations are detailed for individual analyses.
Although we quantified concentrations of several trace elements, we limited our
statistical analyses to those elements commonly reported in the D-area ash basins where
our study was conducted that are a concern to human health and to the health and survival
of organisms occupying ash basins (As, Se, Hg; Luther 2010a,b). Other element levels
are presented for descriptive purposes only. As, Se, and Hg are bioavailable to all trophic
levels in D-area and have potential for antagonistic or synergistic interactions, which can
affect accumulation patterns (Gaines et al. 2002, Hopkins et al. 1999, Unrine et al. 2007).
Interactions among these three elements have been found in laboratory studies of
waterfowl and studies of different organisms in situ in contaminated areas. To determine
tissue-specific accumulation patterns of these elements over the course of our experiment,
we log-transformed concentrations of As, Se, and Hg to improve distributions and
subsequently developed separate linear regression models for each element over time. To
assess potential interactions among elements for naturally exposed waterfowl we utilized
97
a Spearman correlation test among As, Se, and Hg with days of exposure included as a
covariate to control for expected change over time.
To verify the utility of blood samples as a non-destructive sampling technique to
predict element concentrations in muscle and liver in future studies, we first compared 4
common data transformations (inverse, square root, log, and exponential of the
concentrations) with the untransformed concentrations (raw data) and chose the best fit
(R2 value) linear regression model for each tissue relative to post-experiment blood levels
for As and Se (Hodgman et al. 1996). Hg was not included because >50% of the blood
values were BDL. We then ran a k-fold cross validation to confirm our predictive
equations for As and Se would work on novel data (Package DAAG version 1.2). We
choose to evaluate three folds of the data, testing 1/3 of the data at a time against the
model developed for each element. The overall mean sums of squares (MS) for the 3-
folds was compared to the original model residual standard error (RSE) to determine how
closely the MS and RSE aligned to assess whether the equations generated were reliable
for predictions on unseen data.
To assess human consumption limits for ducks utilizing ash basins over time
periods relevant to migratory waterfowl, we binned collections into 6 time periods of
exposure, each ~15 days. The Environmental Protection Agency (EPA) has established
chronic oral reference dose limits for As, Se, and Hg based on consumption of fish;
concentrations a human could theoretically be exposed to daily over the course of their
life and not expect detrimental health consequences. Although we quantified total THg,
which includes all mercury in a given sample, nearly all Hg in tissues of higher trophic
level organisms such as birds and fish is in the methylmercury (MeHg) form; 80-100% in
98
muscle of piscivorous birds and up to 98% in fish muscle (Wiener et al. 2003, Evers et al.
2005). Thus, THg can be assumed to approximate MeHg concentrations for consumption
risk analyses (Cristol et al. 2012). MeHg also is the most bioavailable form for humans
and wildlife to absorb and the most toxic form of mercury, which can cause severe
neurological damage. Therefore, calculations of consumption limits were performed with
EPA limits for MeHg (USEPA 1999a, 2001; Grandjean et al. 1999; Hall et al. 2009;
Clarkson et al. 2003). We used the average and maximum concentrations of the
aforementioned trace elements found in muscle samples to calculate human consumption
limits and provide estimates of the average and worst-case exposure risk scenarios for
consuming waterfowl with exposure in the range of days represented in each exposure
group.
For our calculations of consumption limits of waterfowl muscle, we utilized
established EPA equations for fish advisory limits; including average standard weights
for adults and children (70 kg or 154 lbs and 16 kg or 35 lbs, respectively) and meal sizes
(227g or ½ lb for adults and 113g or ¼ lb for children). When choosing a metric to set
our threshold, we used the EPA established chronic oral reference doses of 0.0001
milligrams per kilograms per day (mg/kg/day) for MeHg (USEPA 1999a), 0.005
mg/kg/day for Se (USEPA 1999b) and 0.0003 mg/kg/day for As (USEPA 1998). These
levels are daily dose exposures that alone are unlikely to produce appreciable deleterious
effects over a lifetime of exposure; as exposures increase above the reference doses so
does the risk of adverse health effects.
Most toxicology studies dealing with tissues report either wet (raw sample) or dry
mass (devoid of water), but neither are representative of how most people consume
99
muscle tissue. Therefore, we calculated trace element concentrations for consumption
limits based on what we would expect in a cooked sample. We utilized the average
percent moisture loss in cooked duck breast muscle (28.2%; Omojola 2007) to amend
concentrations of trace elements that would be found in cooked muscle. This gives a
higher concentration of trace elements than the raw sample and lower concentrations than
the dry sample, but a more accurate assessment of real world scenarios for human
consumption of game meat.
RESULTS
We released 90 ring-necked ducks onto ash basin 1 in D-Area in two release
periods (December 3rd-18th 2014 and February 3rd-10th 2015). Between January 24th and
March 12th 2015 we collected 38 ducks that ranged between 3 and 92 days of exposure
(Table 3-1). Cd and Pb in muscle and blood and Hg in post-experiment blood had >50%
of samples BDL and thus were excluded from presented descriptive statistics (Table 3-2).
Pre-experiment blood values indicated 5 birds exhibited As, Se or Hg levels two standard
deviations above the mean and thus were excluded from analysis.
For muscle, liver, and blood the relationship between log-transformed As, Se, and
Hg concentrations and time of exposure showed Se had a significant positive relationship
with days of exposure for muscle and blood (Table 3-3). As concentrations in muscle,
liver and blood did not differ significantly over the course of our experiment days of
exposure (Table 3-4). Hg had a significant positive relationship with days of exposure in
muscle and a significant negative relationship with days of exposure in liver (Table 3-5).
When assessing potential interactions among elements we found that none of the
elements were significantly correlated in muscle, but in liver Se and As were significantly
100
correlated, as were Se and Hg in liver (Table 3-6). When comparing element levels
between muscle and liver tissues, we observed a significant correlation between tissues
for both As and Se, but not Hg (Table 3-6). Investigation into whether element
concentrations in post-experiment blood were correlated with levels observed in other
tissues revealed that blood concentrations of As were significantly correlated with muscle
and liver. Blood concentrations of Se were significantly correlated with muscle.
Our regression models quantifying the relationship between post-collection blood
element concentrations and those observed in muscle and liver tissue showed that for As
the raw data gave a better fit than any of the transformations we tried for muscle, and for
liver the inverse of the concentration provided the best fit. The regression equation for
muscle was and liver was , both with R2 values
of 0.56. For Se the inverse concentration was the best transformation for muscle, giving
the equation with a R2 of 0.96; none of the transformations
improved the fit beyond the raw concentrations for Se in liver, giving the equation
(R2 of 0.14).
Results of our k-fold cross validation models showed that for As in muscle the
overall mean sums of squares (MS; 0.17) for the 3 folds of the data was less than the
model residual standard errors (RSE; 0.40). This indicated that the equation was reliable
for prediction on unseen data because the error expected on unseen data would have been
close to or below our original error value for the RSE. For liver, the equation for As was
also reliable with the overall MS of 0.54 less than the RSE of 0.66. Overall MS for Se in
muscle (1.06) was very close to the RSE (0.99), and indicates this equation would
produce estimates with about a 7% increase in error expected with novel data. In liver Se
y = 0.5648x +0.3897 y −1 = 0.2872*e −1 +0.2414
y −1 =1.3145* x −1 +0.0272
y = 0.4916x +10.529
101
had a much higher overall MS (57.80) than RSE (7.32) and therefore would not be
sufficient to provide reliable predictions on unseen data.
Moisture loss in muscle from freeze-drying ranged between 66.6-77.9% and our
calculated estimate for moisture content for cooked muscle of 43.0% was used in all
consumption limit calculations (Table 3-7). For As, the number of meals allowed per
month for adults based on average concentrations found in every group ranged from 2.9-
11.8 meals, but varied inconsistently through time, with the lowest levels observed in
Group 2, and the highest in Group 3. Children allowances based on average As
concentrations were between 1.5 and 5.4 meals per month, with 4 out of 6 groups at or
below 1.4 meals per month at the maximum concentrations found. Adult allowances
based on average Se concentrations declined over time from 26.7 meals per month in
Group 1 down to 6.2 meals per month in Group 6; children’s meals allowed also declined
through time from 12.2 meals in Group 1 to 2.8 meals in Group 6. For Hg, adults would
be allowed between 27.5-33.1 meals per month compared to 12.6-15.2 for children at the
average concentrations across all sampling groups.
DISCUSSION
This study provides the most comprehensive assessment of rates of trace element
accumulation in waterfowl utilizing CCW impoundments to date (White et al. 1986). The
results of this study are expected to assist in elucidating potential risks to waterfowl
utilizing coal fly ash settling basins and also human hunters and their families that could
eat these birds. We focused on three common trace elements found in CCW (As, Se, Hg)
that can have deleterious consequences for exposed wildlife and human consumers
(Heinz and Hoffman 1998, Rowe et al. 2002). Of the elements evaluated, only Se clearly
102
accumulated through time, a pattern observed in muscle, liver, and blood tissues.
Surprisingly, Hg concentrations in muscle increased and in liver decreased during the
course of this study, possibly due to interactive effects with Se (Hopkins et al. 2007) as
both As and Hg were correlated with Se in liver tissue.
The accumulation of Se in muscle tissue through time, resulted in a corresponding
decrease in the number of allowable meals per month, decreasing in adults from 26.7
meals per month in our first exposure group to 6.2 in our last exposure group (decreasing
from 12.2 to 2.8 meals per month in children) which had exposures of 76-92 days, a
scenario akin to ducks overwintering in a contaminated area. Despite this decrease, for
adults these estimates still exceed average numbers of waterfowl meals consumed per
month by hunters in the southeastern U.S., but are within the reported range consumed
(Smith et al. Unpublished Manuscript).
Average levels of As in muscle ranged from 0.42-1.70 ppm dw and consumption
limit calculations from these concentrations resulted in adult meal allowances ranging
from 2.9-11.8 meals per month. The European Food Safety Authority (EFSA) advises
that there is no “safe” threshold for As in food, so exposure to As from CCW should be
carefully monitored in game species (EFSA 2009). Although As accumulation in muscle
tissue did not increase over the course of our experiment, As concentrations in waterfowl
should be quantified and sources of As need to be investigated given waterfowl ability to
move extensive distances from point sources. Moreover, our results indicated As had a
negative relationship with Se in muscle tissue, so presence of Se may have ameliorated
accumulation of As, which has been seen previously in laboratory experiments with
mallards (Heinz 1979, Heinz 1996, Hoffman et al. 1992, Stanley et al. 1994).
103
In a previous study (Chapter 2) we opportunistically collected ducks from this
same ash basin (with no known residence times) and reported higher levels of Hg than
were observed in the present study, with average muscle levels of 0.79 ppm dw and
levels ranging from 0.07-5.55 ppm dw. Thus while our current study had high average
meal allowances of 33.1 meals per month for adults, the previous opportunistically
collected ducks from this same basin had average meal allowances of only 2.1 meals per
month for adults and 1.0 meal for children, reducing to 0.3 meals per month for adults
and 0.1 meals for children at maximum observed concentrations (Chapter 2). The
underlying reason for this discrepancy is unknown but likely due to a combination of two
factors. First, as has been seen in laboratory studies for waterfowl, the presence of Se can
ameliorate uptake of Hg concentrations (Heinz 1979, Heinz 1996, Heinz and Hoffman
1998), a pattern also documented in fish and amphibians in CCW contaminated areas
(Southworth et al. 1994, 2000; Hopkins et al. 2006; Peterson et al. 2009; Yang et al.
2010; Metts et al. 2013, Penglase et al. 2014). The significant interaction between Hg and
Se concentrations in muscle tissue further support the likelihood that substantial uptake of
Se limited uptake of Hg in our experiment. Secondly, species sampled in (Chapter 2)
included a variety of diving ducks but no ring-necked ducks and thus low uptake of Hg in
the present study may reflect species-level differences due to diet. Correspondingly, our
results indicated the importance that mensurative studies alone may not give the clearest
possible picture of wildlife exposure risks and thus experimental in situ exposure studies
are needed to elucidate accumulation patterns in free-ranging populations.
For future studies, our data suggest there is a reasonable correlation between
blood As levels and concentrations found within muscle and liver tissues, suggesting
104
blood samples may be used as a non-lethal evaluation tool for quantifying As levels in
free-ranging waterfowl. The relationship between Se in muscle and post-exposure blood
was slightly weaker, with approximately a 7% increase in error expected if the equation
were used on unseen data. The addition of more samples from future studies may be
needed to thoroughly elucidate this relationship. Unfortunately the equation with the best
fit for the relationship of Se liver to post-collection blood did not provide good predictive
value for unseen data.
Taggart et al. (2011) have remarked that game meat may represent a poorly
regulated risk to human health, as game meat can contain levels of toxic elements that
pose a potential risk especially for “at risk” groups like pregnant women, children, or
subsistence hunters (who may consume higher than average amounts of game meat;
Duchesne et al. 2004). Our efforts to connect length of exposure with organ/tissue
burdens for a variety of common trace elements associated with coal ash may be able to
be combined with behavioral data on use of ash basins by free-ranging populations to get
representative estimates of exposure. The data we present on rates of accumulation for
several common CCW contaminants will aid potential monitoring programs to assess
contamination burdens in waterfowl across broader geographical scales to more clearly
elucidate potential risks to wildlife health and human consumers of waterfowl.
MANAGEMENT IMPLICATIONS
Surface impoundments are a potential threat to wildlife because they are often
attractive habitat for many species, which can result in trace element exposure and
increased risk of adverse health effects. We found pollution from coal combustion wastes
has introduced appreciable concentrations of some trace elements into game meat.
105
Although levels were generally below those that likely could result in adverse health
effects to the birds or trigger human consumption advisories, the suggested average daily
Se intake for adult humans (0.06 mg; Rayman 2004) is equal to 1.8 mg per month,
approximately the cooked muscle tissue concentrations found in birds in the first
exposure time period of this experiment. This coupled with the fact that the average daily
intake of Se for women (0.09 mg) and men (0.13 mg; Fairweather-Tait et al. 2011)
already exceeds the suggested intake should warrant at least concern about waterfowl
exposure to Se and monitoring what levels could be passed on to human consumers.
There are few studies of this nature that focus on muscle tissue, rather than liver and
kidney, and because of the mobility of waterfowl the scale of this contamination pathway
is difficult to assess. Our data suggest assessments of trace element concentrations in
blood could be reliable predictors of contaminant burdens in muscle and liver for some
elements. Although the relationship between element accumulation in muscle/liver and
feathers was not evaluated in this study, such a comparison could provide an additional
useful non-lethal metric (or hunter assisted method) for assessing waterfowl exposure to
contaminants. Such methodology could be integrated into a comprehensive study
assessing contaminant exposure patterns across the 3 main flyways, from Canada across
the U.S. and down to Mexico could facilitate long-term monitoring of contaminant
exposure patterns. Evers et al. (2005) conducted a meta-analysis focusing on Hg (or
MeHg) in waterfowl in the Northeastern U.S., but there remains a need for a more
comprehensive and extensive survey of multiple elements of animal and human health
concern to facilitate assessment of potential risks of contaminant exposure.
106
Table 3-1. Tissue and blood concentrations of arsenic (As), selenium (Se), and mercury
(Hg) of each recollected ring-necked duck restricted to the D-Area ash basins (n=33) on
the Savannah River Site (SRS) in the winter of 2014-2015 with between 3 and 92 days of
exposure.
Muscle Liver Blood Id Days Exp. As Se Hg As Se Hg As Se Hg
195 3 0.15 0.72 0.04 1.63 20.79 0.15 0.24 0.90 0 197 3 0.33 0.84 0.02 1.76 20.56 0.25 0.53 1.33 0 189 8 0.80 7.63 0.06 3.69 21.04 0.15 1.42 14.07 0 180 9 1.57 3.19 0.06 4.62 9.42 0.07 2.35 6.26 0 178 19 2.20 4.91 0.03 4.91 15.76 0.12 2.78 14.04 0 188 22 1.84 5.28 0.05 3.41 16.10 0.10 2.18 18.92 0 190 22 1.49 6.03 0.04 2.51 17.21 0.25 2.32 15.81 0.02 179 23 1.84 5.96 0.06 5.13 15.08 0.13 2.29 15.75 0 192 29 1.14 6.77 0.06 1.37 21.24 0.22 0.94 17.72 0.02 110 32 0.25 9.13 0.06 0.47 11.78 0.15 0.17 15.08 0 194 34 1.11 8.69 0.03 1.89 20.34 0.15 1.63 22.08 0.02 171 40 0.22 5.15 0.06 0.35 6.52 0.11 0.16 7.75 0.06 150 43 0.33 8.62 0.04 0.62 9.42 0.09 0.29 12.28 0 148 44 0.20 5.69 0.06 0.27 6.10 0.11 0.21 9.03 0 34 48 0.17 7.04 0.08 0.40 7.29 0.08 0.23 8.54 0.07 15 52 0.35 9.36 0.05 0.84 10.89 0.12 0.41 10.03 0 3 55 0.27 9.46 0.03 0.27 5.64 0.07 0.09 9.64 0 32 56 1.03 10.33 0.06 2.46 25.00 0.11 1.47 14.71 0.04 39 56 0.46 9.56 0.07 1.28 16.86 0.12 0.57 11.96 0 169 64 0.73 13.08 0.05 1.89 30.52 0.07 0.91 20.98 0.09 136 67 1.68 6.88 0.06 1.48 14.88 0.09 2.24 17.17 0.02 144 67 0.83 10.05 0.06 1.70 21.75 0.07 0.90 20.70 0.02 158 67 0.89 8.72 0.03 2.41 14.66 0.11 0.77 14.98 0 160 67 0.34 7.18 0.07 0.70 10.23 0.10 0.35 10.58 0 164 70 1.26 10.54 0.06 2.40 17.95 0.08 1.37 20.70 0.02 95 73 0.99 12.18 0.05 1.98 20.26 0.19 1.01 19.18 0 26 75 0.58 14.33 0.06 1.32 22.52 0.11 0.52 16.64 0 168 78 2.13 11.57 0.07 3.83 24.01 0.11 0.28 7.09 0 41 79 1.26 13.46 0.05 2.42 18.16 0.07 1.18 20.41 0 43 79 1.31 11.92 0.05 3.00 26.06 0.19 0.21 9.61 0 97 85 1.06 12.32 0.06 2.48 17.99 0.07 1.19 19.60 0.02 85 90 1.03 15.64 0.08 2.57 38.99 0.11 0.15 20.01 0 36 92 1.20 15.03 0.06 1.00 7.29 0.07 0.84 25.41 0
107
Table 3-2. Trace elements in ring-necked ducks (n=33) before and after restriction to the D-Area ash basins on the Savannah
River Site (SRS) for between 3 and 92 days in winter of 2014-2015. Concentration mean±SE at ~15-day exposure increments.
Element Tissue 3-15 days (n=4) 16-30 days (n=5) 31-45 days (n=5) 46-60 days (n=5) 61-75 days (n=8) 76-92 days (n=6) V Muscle 0.05±0.00 0.07±0.01 0.08±0.01 0.09±0.01 0.08±0.01 0.10±0.02
Liver 0.35±0.04 0.34±0.07 0.28±0.04 0.42±0.10 0.41±0.04 0.42±0.06
Blood 0.06±0.01 0.05±0.00 0.13±0.04 0.08±0.01 0.08±0.02 0.10±0.02
! Cr Muscle 0.70±0.05 0.74±0.06 0.73±0.04 0.75±0.05 0.76±0.02 0.85±0.06
Liver 2.22±1.11 1.00±0.08 1.03±0.10 1.14±0.14 1.34±0.07 1.22±0.16
Blood 1.12±0.15 1.06±0.05 1.28±0.12 1.05±0.11 1.09±0.06 1.48±0.17
! Ni Muscle 0.05±0.00 0.06±0.02 0.07±0.01 0.10±0.02 0.07±0.01 0.11±0.02
Liver 4.59±4.29 0.39±0.09 0.14±0.06 0.13±0.05 0.19±0.04 0.28±0.03
Blood 0.18±0.04 0.20±0.01 0.22±0.04 0.19±0.03 0.17±0.01 0.30±0.06
! Cu Muscle 34.12±1.93 34.38±1.73 32.09±3.27 29.94±1.86 33.99±1.91 45.41±6.03
Liver 346.20±43.13 286.93±54.47 227.06±42.70 262.44±56.75 288.42±46.88 184.60±70.66
Blood 2.26±0.08 2.41±0.13 2.16±0.26 2.10±0.36 2.67±0.34 2.60±0.36
! Zn Muscle 36.15±2.94 37.86±1.95 34.32±2.69 32.03±1.10 35.85±1.65 43.25±2.76
Liver 171.05±28.20 178.53±8.19 135.38±8.93 147.13±21.62 158.92±5.49 152.29±22.26
Blood 18.46±0.44 20.25±0.87 20.59±0.82 17.43±1.18 19.70±0.72 19.07±0.84
! As Muscle 0.71±0.32 1.70±0.18 0.42±0.17 0.45±0.15 0.91±0.15 1.33±0.17
Liver 2.92±0.74 3.47±0.71 0.72±0.30 1.05±0.40 1.73±0.20 2.55±0.38
Blood 1.13±0.48 2.10±0.31 0.49±0.29 0.55±0.24 1.01±0.21 0.64±0.20
! Se Muscle 3.10±1.61 5.79±0.32 7.45±0.84 9.15±0.56 10.37±0.96 13.32±0.69
Liver 17.95±2.85 17.08±1.10 10.83±2.59 13.14±3.54 19.09±2.19 22.08±4.31
Blood 5.64±3.06 16.45±0.85 13.25±2.55 10.98±1.09 17.62±1.27 17.02±2.89
! Hg Muscle 0.05±0.01 0.05±0.01 0.05±0.01 0.06±0.01 0.05±0.00 0.06±0.00
Liver 0.16±0.04 0.17±0.03 0.12±0.01 0.10±0.01 0.10±0.01 0.10±0.02
!
108
Cd Liver 1.16±0.48 1.26±0.10 0.92±0.31 0.83±0.16 1.14±0.22 1.24±0.17
! Pb Liver 0.51±0.35 1.07±0.87 0.21±0.10 0.12±0.05 0.08±0.03 0.26±0.09 a Specific days of exposure within groups were (3-15 days, 19-29 days, 31-44 days, 48-56 days, 64-75 days, 78-92 days)
b For Hg in blood 22 (66%) were BDL
c For Cd in muscle 28 (85%), and all blood samples, were BDL
d For Pb in muscle 30 (91%) and in blood 27 (82%) were BDL
109
Table 3-3. Selenium (Se) linear regression with days of exposure for recollected ring-necked ducks restricted to the D-Area
ash basins (n=33) between 3 and 92 days on the Savannah River Site (SRS) in the winter of 2014-2015.
Muscle Liver Blood Source Estimate SE t P Estimate SE t P Estimate SE t P
As Intercept -0.60 0.30 -1.99 0.05 0.58 0.32 1.83 0.08 -0.03 0.37 -0.08 0.93 DaysExp 0.01 0.01 1.05 0.30 0.00 0.00 -0.50 0.62 -0.01 0.01 -1.21 0.24
Table 3-4. Arsenic (As) linear regression with days of exposure for recollected ring-necked ducks restricted to the D-Area ash
basins (n=33) between 3 and 92 days on the Savannah River Site (SRS) in the winter of 2014-2015.
Muscle Liver Blood Source Estimate SE t P Estimate SE t P Estimate SE t P
Se Intercept 1.00 0.17 6.02 <0.0001 2.61 0.18 14.20 <0.0001 1.79 0.24 7.60 <0.0001 DaysExp 0.02 0.00 6.91 <0.0001 0.00 0.00 0.75 0.46 0.01 0.00 3.36 0.002
Table 3-5. Mercury (Hg) linear regression with days of exposure for recollected ring-necked ducks restricted to the D-Area
ash basins (n=33) between 3 and 92 days on the Savannah River Site (SRS) in the winter of 2014-2015. Blood Hg
concentrations had >50% BDL and thus were not tested statistically.
Muscle Liver Source Estimate SE t P Estimate SE t P
Hg Intercept -3.17 0.12 -29.63 <0.0001 -1.86 0.13 -14.13 <0.0001 DaysExp 0.004 0.00 2.27 0.03 -0.01 0.00 -2.75 0.01
110
Table 3-6. Correlationsa among trace element concentrations in muscle (above diagonal)
and liver (below diagonal) tissues for recollected ring-necked ducks restricted to the D-
Area ash basins (n=33) between 3 and 92 days on the Savannah River Site (SRS) in the
winter of 2014-2015. Correlations between muscle and liver samples for individual
elements are presented on the diagonal in bold.
As Se Hg As 0.81 (<0.0001) NS NS Se 0.53 (<0.001) 0.46 (0.005) NS Hg NS 0.35 (0.04) NS
aSpearman correlation coefficients and P values in parentheses; NS=not significant (P >
0.05)
111
Table 3-7. The monthly allowances of ½ lb. meals for adults and ¼ lb. for children before exceeding the EPA’s chronic oral
reference dose limits for arsenic (As), selenium (Se), and mercury (Hg) for muscle tissue of ring-necked ducks collected from
the D-Area ash basins on the Savannah River Site (SRS) after being restricted between 3 and 92 days of exposure a.
Consumption limits based on average concentrations of cooked ducks muscle are presented with limits based on the maximum
concentration found in an individual for each trace element in parentheses.
Element Group 3-15 days
(n=4) 16-30 days
(n=5) 31-45 days
(n=5) 46-60 days
(n=5) 61-75 days
(n=8) 76-92 days
(n=6) As Adult 7.0 (3.2) 2.9 (2.3) 11.8 (4.5) 11.0 (4.8) 5.4 (3.0) 3.7 (2.3)
Child 3.2 (1.4) 1.3 (1.0) 5.4(2.1) 5.1 (2.2) 2.5 (1.4) 1.7 (1.1)
! ! !Se Adult 26.7 (10.8) 14.3 (12.2) 11.1 (9.1) 9.0 (8.0) 8.0 (5.8) 6.2 (5.3)
Child 12.2 (5.0) 6.6 (5.6) 5.1 (4.2) 4.1 (3.7) 3.7 (2.6) 2.8 (2.4)
! ! !Hg Adult 33.1 (33.1) 33.1 (33.1) 33.1 (33.1) 27.5 (20.7) 33.1 (23.6) 27.5 (20.7) !! Child 15.2 (15.2) 15.2 (15.2) 15.2 (15.2) 12.6 (9.5) 15.2 (10.8) 12.6 (9.5)
a Specific days of exposure within groups were (3-15 days, 19-29 days, 31-44 days, 48-56 days, 64-75 days, 78-92 days)
112
Figure 3-1. D-Area ash basins on the Savannah River Site (SRS), SC. Basin 1, the largest
basin is partially filled in and has extensively revegetated. The smaller enclosed wetland
formed by revegetation in this basin was utilized as the release and exposure area for the
ring-necked ducks in this study in winter of 2014-2015.
113
Figure 3-2. Muscle concentrations of arsenic (As), selenium (Se), mercury (Hg), and
days of exposure of recollected ring-necked duck restricted to the D-Area ash basins
(n=33) on the Savannah River Site (SRS) in the winter of 2014-2015 between 3 and 92
days of exposure.
R² = 0.003
R² = 0.76
R² = 0.14 0.0 2.0 4.0 6.0 8.0
10.0 12.0 14.0 16.0 18.0
0 15 30 45 60 75 90
Con
cent
ratio
n (p
pm d
w)
Days of Exposure
Muscle Tissue
As
Se
Hg
114
Figure 3-3. Liver concentrations of arsenic (As), selenium (Se), mercury (Hg), and days
of exposure of recollected ring-necked duck restricted to the D-Area ash basins (n=33) on
the Savannah River Site (SRS) in the winter of 2014-2015 between 3 and 92 days of
exposure.
R² = 0.05
R² = 0.05
R² = 0.19 0.0 5.0
10.0 15.0 20.0 25.0 30.0 35.0 40.0 45.0
0 15 30 45 60 75 90
Con
cent
ratio
n (p
pm d
w)
Days of Exposure
Liver Tissue
As
Se
Hg
115
Figure 3-4. Blood concentrations of arsenic (As), selenium (Se), mercury (Hg), and days
of exposure of recollected ring-necked duck restricted to the D-Area ash basins (n=33) on
the Savannah River Site (SRS) in the winter of 2014-2015 between 3 and 92 days of
exposure.
R² = 0.11
R² = 0.24
R² = 0.0002 0.0
5.0
10.0
15.0
20.0
25.0
30.0
0 15 30 45 60 75 90
Con
cent
ratio
n (p
pm d
w)
Days of Exposure
Blood Levels Over Time
As
Se
Hg
116
LITERATURE CITED
Beck, M. L., W.A. Hopkins, and B. P. Jackson. 2013. Spatial and temporal variation in
the diet of tree swallows: implications for trace-element exposure after habitat
remediation. Archives of Environmental Contamination and Toxicology 65:575-
587.
Beck, M. L., W. A. Hopkins, B. P. Jackson, and D. M. Hawley. 2015. The effects of a
remediated fly ash spill and weather conditions on reproductive success and
offspring development in tree swallows. Environmental Monitoring and
Assessment 187:1-25.
Boening, D. W. 2000. Ecological effects, transport, and fate of mercury: A general
review. Chemosphere 40:1335-1351.
Bryan Jr., A. L., W. A. Hopkins, J. H. Parikh, B. P. Jackson, and J. M. Unrine. 2012.
Coal fly ash basins as an attractive nuisance to birds: Parental provisioning
exposes nestlings to harmful trace elements. Environmental Pollution 161:170-
177.
Cherry, D. S., S. R. Larrick, R. K. Guthrie, E. M. Davis, and F. F. Sherberger. 1979.
Recovery of invertebrate and vertebrate populations in a coal ash stressed
drainage system. Journal of the Fisheries Research Board of Canada 36:1089-
1096.
Clarkson, T. W., L. Magos, and G. J. Meyers. 2003. Human exposure to mercury: The
three modern dilemmas. Journal of Trace Elements in Experimental Medicine
16:321-343.
117
Coğun, H. Y., T. A. Yüzereroğlu, Ö. Firat, G. Gök, and F. Kargin. 2006. Metal
concentrations in fish species from the Northeast Mediterranean Sea.
Environmental Monitoring and Assessment 121:431-438.
Conder, J. M., and J. A. Arblaster. 2016. Development and use of wild game
consumption rates in human health risk assessments. Human and Ecological Risk
Assessment 22:251-264.
Cristol, D. A., L. Savoy, D. C. Evers, C. Perkins, R. Taylor, and C. W. Varian-Ramos.
2012. Mercury in waterfowl from a contaminated river in Virginia. Journal of
Wildlife Management 76:1617-1624.
DeVault, T. L., B. D. Reinhardt, I. L. Brisbin, Jr., and O. E. Rhodes, Jr. 2004. Home
ranges of sympatric black and turkey vultures in South Carolina. Condor 106:706-
711.
Dorman, L., J. H. Rodgers Jr., and J. W. Castle. 2010. Characterization of ash-basin
waters from a risk-based perspective. Water, Air, and Soil Pollution 206:175-185.
Duchesne, J. F., B. Levesque, D. Gauvin, B. Braune, S. Gingras, and E. Dewailly. 2004.
Estimating the mercury exposure dose in a population of migratory bird hunters in
the St. Lawrence River region, Quebec, Canada. Environmental Research 95:207-
217.
European Food Safety Authority (EFSA). 2009. Panel on contaminants in the food chain
(CONTAM); scientific opinion on arsenic in food, Parma, Italy. EFSA Journal
7:1351-1550.
Evers, D. C., N. M. Burgess, L. Champoux, B. Hoskins, A. Major, W. M. Goodale, R. J.
Taylor, R. Poppenga, T. Daigle. 2005. Patterns and interpretation of mercury
118
exposure in freshwater avian communities in northeastern North America.
Ecotoxicology 14:193-221.
Fairweather-Tait, S. J., Y. Bao, M. R. Broadley, R. Collings, D. Ford, J. E. Hesketh, and
R. Hurst. 2011. Selenium in human health and disease. Antioxidants and Redox
Signaling 14:1337-1383.
Farkas, A., J. Salánki, and A. Specziár. 2003. Age- and size-specific patterns of heavy
metals in the organs of freshwater fish Abramis brama L. populating low-
contaminated site. Water Research 37:959-964.
Fletcher, D. E., A. H. Lindell, G. K. Stillings, G. L. Mills, S. A. Blas, J. V. McArthur.
2014. Variation in trace-element accumulation in predatory fishes from a stream
contaminated by coal combustion waste. Archives of Environmental
Contamination and Toxicology 66:341-360.
Gaines, K. F., C. S. Romanek, C. S. Boring, C. G. Lord, M. Gochfeld, and J. Burger.
2002. Using raccoons as an indicator species for metal accumulation across
trophic levels: A stable isotope approach. Journal of Wildlife Management.
66:811-821.
Grandjean, P., R. F. White, A. Nielsen, D. Cleary, and E. C. Santos. 1999.
Methylmercury neurotoxicity in Amazonian children downstream from gold
mining. Environmental Health Perspectives 107:587-591.
Greeley, M. S., S. M. Adams, L. R. Elmore, and M. K. McCracken. 2016. Influence of
metal(loid) bioaccumulation and maternal transfer on embryo-larval development
in fish exposed to a major coal ash spill. Aquatic Toxicology 173:165-177.
119
Hall, B. D., L. A. Baron, and C. M. Somers. 2009. Mercury concentrations in surface
water and harvested waterfowl from the prairie pothole region of Saskatchewan.
Environmental Science and Technology, 43:8759-8766.
Halverson, N. V., L. D. Wike, K. K. Patterson, J. A. Bowers, A.L. Bryan, K. F. Chen et
al. 1997. SRS environmental information document−SRS ecology chapter
5−streams, reservoirs, and the Savannah River. Section 5.2−Beaver dam creek
drainage description and surface hydrology. WSRC-TR-97-0223, pp. 67-119.
Havelková, M., L. Dušek, D. Némethová, G. Poleszczuk, and Z. Svobodová 2008.
Comparison of mercury between liver and muscle−A biomonitoring of fish from
lightly and heavily contaminated localities. Sensors 8:4095-4109.
Heinz, G. H. 1979. Methylmercury: Reproductive and behavioral effects on three
generations of mallard ducks. Journal of Wildlife Management 43:394-401.
Heinz, G. H. 1996. Selenium in birds. Environmental Contaminants in Wildlife:
Interpreting Tissue Concentrations. (Eds. W. N. Beyer, G.H. Heinz, and A. W.
Redmon-Norwood) CRC Press, Boca Raton, FL, pp. 447-458.
Heinz, G. H., and D. J. Hoffman. 1998. Methylmercury chloride and selenomethionine
interactions on health and reproduction in mallards. Environmental Toxicology
and Chemistry 17:139-145.
Hodgman, T. P., B. B. Davitt, and J. R. Nelson. 1996. Monitoring mule deer diet quality
and intake with fecal indices. Journal of Range Management 49:215-222.
Hoffman, D.J., C. J. Sanderson, L. J. LeCaptain, E. Cromartie, and G. S. Pendleton. 1992.
Interactive effects of arsenic, selenium, and dietary protein on survival, growth,
120
and physiology in mallard ducklings. Archives of Environmental Contamination
and Toxicology 20:288-294.
Hopkins, W. A., C. L. Rowe, and J. D. Congdon. 1999. Elevated trace element
concentrations and standard metabolic rate in banded water snakes (Nerodia
fasciata) exposed to coal combustion wastes. Environmental Toxicology and
Chemistry 18:1258-1263.
Hopkins, W. A., J. D. Congdon, and J. K. Ray. 2000. Incidence and impact of axial
malformations in larval bullfrogs (Rana catesbeiana) developing in sites polluted
by a coal-burning power plant. Environmental Toxicology and Chemistry 19:862-
868.
Hopkins, W. A., S.E. DuRant, B. P. Staub, C. L. Rowe, and B. P. Jackson. 2006.
Reproduction, embryonic development, and maternal transfer of contaminants in
the amphibian Gastrophryne carolinensis. Environmental Health Perspectives
114:661-666.
Hopkins, W. A., L. B. Hopkins, J. M. Unrine, J. Snodgrass, and J. D. Elliot. 2007.
Mercury concentrations in tissues of osprey from the Carolinas, USA. The Journal
of Wildlife Management 71:1819-1829.
Hoppe, R. T., L. M. Smith, and D.B. Webster. 1986. Foods of wintering diving ducks in
South Carolina. Journal of Field Ornithology 57:126-134.
Horton, J. H. 1974. Technical Division, Savannah River Laboratory. Memorandum to
B.C. Rusche. Subject: Mercury in the separations areas seepage basins. DPST-74-
231. SRS Phase II Database GKH.
121
Humphries, M. 1999. Global climate change: Coal use in China and other Asian
developing countries. Washington, D.C. Congressional Research Service, the
Library of Congress. University of North Texas Digital Library
http://digital.library.unt.edu/ark:/67531/metacrs848/. Accessed 11 January 2016.
Ikem, A., N. O. Egiebor, and K. Nyavor. 2003. Trace elements in water, fish, and
sediment from Tuskegee Lake Southeastern USA. Water, Air, and Soil Pollution
149:51-75.
Kennamer, R. A. 2003. Recoveries of ring-necked ducks banded on the U.S. Department
of Energy’s Savannah River Site, South Carolina. The Oriole 68:8-14.
Kennamer, R. A. 2005. Waterfowl. Ecology and management of a forested landscape:
Fifty years on the Savannah River Site. (Eds. J. C. Kilgo and J. I. Blake) Island
Press, Washington, D.C., pp. 347-359.
Lemly, A. D. 1996. Selenium in aquatic organisms. Environmental Contaminants in
Wildlife: Interpreting Tissue Concentrations. (Eds. W. N. Beyer, G. H. Heinz, and
A. W. Redmon-Norwood) CRC Press, Boca Raton, FL, pp. 427-455.
Lemly, A. D. 2002. Symptoms and implications of selenium toxicity in fish: the Belews
Lake case example. Aquatic Toxicity 57:39-49.
Lemly, A. D., and J. Skorupa. 2012. Wildlife and coal waste policy debate: Proposed
rules for coal waste disposal ignore lessons from 45 years of wildlife poisoning.
Environmental Science and Technology 46:8595-8600.
Lide, R. F. 1994. Carolina bays and similar natural depression wetlands of the SRS. GIS
coverage in ARC/INFO. Savannah River Ecology Laboratory, Savannah River
Site, Aiken, SC.
122
Luther, L. 2010a. Managing coal combustion waste (CCW): Issues with disposal and use.
Congressional Research Service. CRS report for Congress, 7-5700, R40544, pp.
1-26. https://www.fas.org/sgp/crs/misc/R40544.pdf. Accessed 29 January 2016.
Luther, L. 2010b. Regulating Coal Combustion Waste Disposal: Issues for Congress.
Congressional Research Service. CRS report for Congress, 7-5700, R41341, pp.
1-22. http://www.fas.org/sgp/crs/misc/R41341.pdf. Accessed 29 January 2016.
Mason, R. P., J-M Laporte, and S. Andres. 2000. Factors controlling the bioaccumulation
of mercury, methylmercury, arsenic, selenium, and cadmium by freshwater
invertebrates and fish. Archives of Environmental Contamination and Toxicology
38:283-297.
Mayer, J. J., R. A. Kennamer, and R. T. Hoppe. 1986. Waterfowl of the Savannah River
Plant. Final Report. Report SREL-22 UC-66e. Savannah River Ecology
Laboratory, Division of Stress and Wildlife Ecology, Aiken, SC.
Mayfield, D. B., S. Thakali, W. T. Mehler, A. S. Lewis. 2013. Ecological effects of coal
combustion products (CCPs). A literature review of observed effects and
considerations for managing risks. 2013 World of Coal Ash Conference,
Lexington, KY. http://www.flyash.info/. Accessed 22 February 2016.
Metts, B. S., K. A. Buhlmann, T. D. Tuberville, D. E. Scott, and W. A. Hopkins. 2013.
Maternal transfer of contaminants and reduced reproductive success of southern
toads (Bufo [Anaxyrus] terrestris) exposed to coal combustion waste.
Environmental Science and Technology 47:2846-2853.
Omojola, A. B. 2007. Carcass and organoleptic characteristics of duck meat as influenced
by breed and sex. International Journal of Poultry Science 6:329-334.
123
Otter, R. R., F. C. Bailey, A. M. Fortner, and S. M. Adams. 2012. Trophic status and
metal bioaccumulation differences in multiple fish species exposed to coal ash-
associated metals. Ecotoxicology and Environmental Safety 85:30-36.
Penglase, S., K. Hamre, and S. Elligen. 2014. Selenium and mercury have a synergistic
negative effect on fish reproduction. Aquatic Toxicology Journal 149:16-24.
Peterson, S. A., N. V. Ralston, D. V. Peck, J. V. Sickle, J. D. Robertson, V. L. Spate, and
J. S. Morris. 2009. How might selenium moderate the toxic effects of mercury in
stream fish of the western US? Environmental Science and Technology 43:3919-
3925.
Rayman, M. P. 2004. The use of high-selenium yeast to raise selenium status: How does
it measure up? British Journal of Nutrition, 92:557-573.
Reash, R. J. 2012. Selenium, arsenic, and mercury in fish inhabiting a fly ash exposure
gradient: Interspecific bioaccumulation patterns and elemental associations.
Environmental Toxicology and Chemistry 31:739-747.
Rice, K. M., E. M. Walker Jr., M. Wu, C. Gillette, and E. R. Blough. 2014.
Environmental mercury and its toxic effects. Journal of Preventative Medicine
and Public Health 47:74-83.
Rigg, D. K., M. N. Wacksman, J. Iannuzzi, T. F. Baker, M. Adams, and M. S. Greeley Jr.
2015. Assessing ecological risks to the fish community from residual coal fly ash
in Watts Bar Reservoir, Tennessee. Integrated Environmental Assessment and
Management 11:88-101.
124
Roe, J. H., W. A. Hopkins, S. E. DuRant, and J. M. Unrine. 2006. Effects of competition
and coal-combustion wastes on recruitment and life history characteristics of
salamanders in temporary wetlands. Aquatic Toxicology 79:176-184.
Rowe, C. L., O. M. Kinney, A. P. Fiori, and J. D. Congdon. 1996. Oral deformities in
tadpoles (Rana catsbeiana) associated with coal ash deposition: Effects on
grazing ability and growth. Freshwater Biology 36:723-730.
Rowe, C. L., W. A. Hopkins, J. D. Congdon. 2002. Ecotoxicological implications of
aquatic disposal of coal combustion residues in the United States: A review.
Environmental Monitoring and Assessment 80:207-276.
Ruhl, L., A. Vengosh, G. S. Dwyer, H. Hsu-Kim, A. Deonarine, M. Bergin, and J.
Kravchenko. 2009. Survey of environmental and health impacts in the immediate
aftermath of the coal ash spill in Kingston, Tennessee. Environmental Science and
Technology 43:6326-6333.
Ruhl, L., A. Vengosh, G. S. Dwyer, H. Hsu-Kim, G. Schwartz, A. Romanski, and S. D.
Smith. 2012. The impact of coal combustion residue effluent on water resources:
A North Carolina example. Environmental Science and Technology 46:12226-
12233.
Savannah River Nuclear Solutions, LLC. 2011. Environmental, safety, health, and quality
regulatory integration and environmental services. Environmental Management
System (EMS) description manual.
http://www.srs.gov/general/pubs/envbul/documents/ems_manual.pdf Accessed
15 October 2014.
125
Scheuhammer, A. M. 1987. The chronic toxicity of aluminum, cadmium, mercury and
lead in birds. Environmental Pollution 46:263-295.
Smith, J., T. Tuberville, and J. C. Beasley. Hunting and consumption patterns in South
Carolina and Georgia hunters. Unpublished manuscript.
Southworth, G. R., M. J. Peterson, and R. R. Turner. 1994. Changes in concentrations of
selenium and mercury in largemouth bass following elimination of fly ash
discharge to a quarry. Chemosphere 29:71-79.
Southworth, G. R., M. J. Peterson, and M. G. Ryon. 2000. Long-term increased
bioaccumulation of mercury in largemouth bass follows reduction of waterborne
selenium. Chemosphere 41:1101-1105.
Stanley Jr., T. R., J. W. Spann, G. J. Smith, and R. Rosscoe. 1994. Main and interactive
effects of arsenic and selenium on mallard reproduction and duckling growth and
survival. Archives of Environmental Contamination and Toxicology 26:444-451.
Stepanauskas, R., T. C. Glenn, C. H. Jagoe, R. C. Tuckfield, A. H. Lindell, and J. V.
McArthur. 2005. Elevated microbial tolerance to metals and antibiotics in metal-
contaminated industrial environments. Environmental Science and Technology
39:3671-3678.
Taggart, M. A., M. M. Reglero, P. R. Camarero, R. Mateo. 2011. Should legislation
regarding maximum Pb and Cd levels in human food also cover large game meat?
Environmental International 37:18-25.
United States Department of Energy (USDOE). 2012.
http://energy.gov/em/articles/savannah-river-site-retires-coal-fired-d-area-
powerhouse-after-nearly. Accessed 8 September 2015.
126
United States Environmental Protection Agency (USEPA). 1998. Integrated risk
information system (IRIS) on arsenic. National Center for Environmental
Assessment, Office of Research and Development, Washington, DC.
http://www.epa.gov/ttnatw01/hlthef/arsenic.html#ref5. Accessed 30 January
2015.
United States Environmental Protection Agency (USEPA). 1999a. Integrated risk
information system (IRIS) on methylmercury. National Center for Environmental
Assessment, Office of Research and Development, Washington, DC.
http://www.epa.gov/ttnatw01/hlthef/mercury.html#ref13. Accessed 30 January
2015.
United States Environmental Protection Agency (USEPA). 1999b. Integrated risk
information system (IRIS) on selenium and compounds. National Center for
Environmental Assessment, Office of Research and Development, Washington,
DC. http://www.epa.gov/airtoxics/hlthef/selenium.html#ref4. Accessed 30
January 2015.
United States Environmental Protection Agency (USEPA). 2001. Water quality criterion
for the protection of human health: Methylmercury. Office of Science and
Technology, Office of Water, Washington DC.
http://water.epa.gov/scitech/swguidance/standards/criteria/health/upload/2009_01
_15_criteria_methylmercury_mercury-criterion.pdf Accessed 15 January 2015.
United States Environmental Protection Agency (USEPA). 2012b. Waste and cleanup
risk assessment glossary. U.S. Environmental Protection Agency Office of Solid
Waste Emergency Response, Washington, DC.
127
http://www.epa.gov/oswer/riskassessment/glossary.html. Accessed 16 October
2014.
United States Environmental Protection Agency (USEPA). 2015. Frequent questions
about coal ash disposal rule. U.S. Environmental Protection Agency, Washington,
DC. http://www.epa.gov/coalash/frequent-questions-about-coal-ash-disposal-
rule#3. Accessed 07 January 2016.
United States Fish and Wildlife Service (USFWS) and United States Census Bureau
(USCB). 2011. 2011 National survey of fishing, hunting and wildlife associated
recreation. U.S. Department of the Interior, Washington, DC.
https://www.census.gov/prod/2012pubs/fhw11-nat.pdf. Accessed 09 February
2015.
Unrine, J. M., W. A. Hopkins, C. S. Romanek, and B. P. Jackson. 2007.
Bioaccumulation of trace elements in omnivorous amphibian larvae: Implications
for amphibian health and contaminant transport. Environmental Pollution
149:182-192.
Van Dyke, J. U., W. A. Hopkins, and B. P. Jackson. 2013. Influence of relative trophic
position and carbon source on selenium bioaccumulation in turtles from a coal
fly-ash spill site. Environmental Pollution 182:45-52.
Wiener, J. G., D. P. Krabbenhoft, G. H. Heinz, A. M. Scheuhammer. 2003.
Ecotoxicology of mercury. Handbook of Ecotoxicology, Second edition. (Eds.
D.J. Hoffman, B.A. Rattner, G.A. Burton Jr., J. Cairns Jr.) CRC Press, Boca
Raton, FL, pp 409-463.
128
White, D. H., K. A. King, C. A. Mitchell, and B. M. Mulhern. 1986. Trace elements in
sediments, water, and American coots (Fulica americana) at a coal-fired power
plant in Texas, 1979–1982. Bulletin of Environmental Contamination and
Toxicology, 36:376-383.
White, D. L., and K. E. Gaines. 2000. The Savannah River Site: Site description, land
use, and management history. Studies in Avian Biology 21:8-17.
Wright, M. S., G. L. Peltier, R. Stepanauskas, and J. V. McArthur. 2006. Bacterial
tolerance to metals and anti-biotic in metal-contaminated and reference streams.
FEMS Microbiology Ecology 58:293-302.
Yang, D.-Y., Y.-W., J. M. Gunn, and N. Betzille. 2010. Inverse relationships between
selenium and mercury in tissues of young walleye (Stizosedion vitreum) from
Canadian boreal lakes. Science of the Total Environment 408:1676-1683.
Yudovich, Y. E., and M. P. Ketris. 2005a. Arsenic in coal: A review. International
Journal of Coal Geology 61:141-196.
Yudovich, Y. E., and M. P. Ketris. 2005b. Mercury in coal: A review. Part 1.
Geochemistry. International Journal of Coal Geology 62:107-134.
Yudovich, Y. E., and M. P. Ketris. 2006. Selenium in coal: A review. International
Journal of Coal Geology 67:112-126.
129
CHAPTER 4
RADIOCESIUM IN WATERFOWL/WATERBIRDS FROM A RETIRED NUCLEAR
REACTOR COOLING RESERVOIR 1
____________________
1 Oldenkamp, R. E., R. A. Kennamer, A. L. Bryan, Jr., and J. C. Beasley. To be submitted to the Journal of Radioactivity.
130
ABSTRACT
Low-levels of radiocesium contamination is distributed throughout much of the
globe due to nuclear accidents, nuclear weapons testing, and releases from nuclear
facilities. Radiocesium can be accumulated in the bodies of wildlife species, particularly
in skeletal muscle. Thus gamebirds, including waterfowl and water birds, can be a
conduit of human exposure to radiocesium, even from closed access contaminated sites
when birds travel hundreds to thousands of km from the contaminated sources. We
released American coots (Fulica americana) and ring-necked (Aythya collaris) ducks
onto a radiocesium contaminated reservoir on the Savannah River Site in Aiken, SC,
USA, then recollected birds after 33-173 days of exposure. We quantified uptake of
radiocesium over time and compared uptake between species in whole-body relationships
as well as concentrations in muscle and liver tissue. For American coots we also
calculated the ecological half-life for radiocesium in our study system by comparing
whole-body burdens to historical data collected from the same location. Coots maintained
equilibrium whole-body radiocesium concentrations from the first collection event at 32
days through 90 days of exposure, while concentrations in ring-necked ducks continued
to increase between 33 and 90 days of exposure. Beyond 90 days tissue concentrations
declined in both species. Ring-necked ducks had higher concentrations of radiocesium
levels in muscle tissue (t = 2.64, P = 0.01), but liver levels were not significantly
different between species. Ninety-seven percent (34 of 35) of coots and 92% (33 of 36) of
ring-necked ducks had muscle concentrations above the European Economic Community
(EEC 1986) limit of 0.600 Bq/g in fresh meat established for human consumption. We
estimated the ecological half-life of radiocesium in American coots at our study site to be
131
~17 years (with a possible range between approximately 13-24 years). Our data suggest
waterfowl/waterbirds represent a possible vector for human exposure to radionuclides
from contaminated areas, especially habitats where birds may stop-over on migration or
areas in which they over-winter.
INTRODUCTION
Releases of radiocesium into the environment can have wide-ranging
consequences for ecosystems and potentially human health. Although low levels of
radiocesium contamination are distributed throughout much of the globe due to nuclear
accidents and nuclear weapons testing, releases from nuclear facilities also have
occurred, often with inputs directly into aquatic ecosystems. Radiocesium is of biological
importance because its chemical behavior is similar to potassium and can be accumulated
in the bodies of wildlife species, particularly in skeletal muscle (Potter et al. 1989).
Moreover, 137Cs, an important isotope of radiocesium, has a relatively long physical half-
life of 30.2 years and thus can remain in aquatic and terrestrial ecosystems for prolonged
periods (Evans et al. 1983).
Gamebirds, including waterfowl and water birds can be a conduit of human
exposure to radiocesium, even from closed-access contaminated sites. Birds that forage in
restricted contaminated areas may be harvested by hunters hundreds to thousands of kms
from the contaminated sources (Kennamer 2003, Cristol et al. 2012). The Savannah River
Site (SRS), a U.S. Department of Energy (DOE) former nuclear production site in South
Carolina has abundant wetlands and large abandoned cooling reservoirs. These resources
are protected from public disturbance and provide important inland breeding, stop-over,
and overwintering habitat for thousands of resident and migratory waterfowl/waterbirds
132
representing as many as 28 species (Kennamer 2005). Many species use the SRS as a
migratory stopover or overwintering location, including ring-necked ducks (Aythya
collaris) that have later been harvested throughout much of eastern North America, as far
north as Manitoba to Nova Scotia, and as far south as Cuba (Kennamer 2003).
Multiple low-level releases of radiocesium have occurred since the SRS was
established in the 1950s (Carlton et al. 1992, 1994), with radiocesium released into
cooling waters from four of the five nuclear reactors (Paller et al. 1999). During the past
three decades principles of radionuclide distribution and decline in SRS aquatic habitats
have been studied in sediments (Mohler et al.1997) and numerous biota such as wood
ducks (Aix sponsa; Fendley et al. 1977), American coots (Brisbin and Vargo 1982,
Brisbin 1991a, Brisbin and Kennamer 2000), snakes (Bagshaw and Brisbin 1984), and
fish (Paller et al. 1999, Peles et al. 2000).. In particular, Pond B on the SRS has been an
important study system over the last several decades as it received the greatest inputs of
radiocesium of the SRS reservoirs (Brisbin 1991a). Previous work in Pond B determined
American coots (Fulica americana) consistently accumulated levels of radiocesium
exceeding other wintering species investigated, including carnivorous, omnivorous, and
piscivorous species (Brisbin et al. 1973). Thus, coots have been used as a model species
for ecological risk assessments in this system (Brisbin et al 1973, Brisbin 1993). Diet is a
likely biological explanation given for this differentiation in accumulation, as coots
primarily consume algae and aquatic vascular plants that are known to accumulate high
levels of radiocesium (Potter 1987, Brisbin et al. 2002). However, these data were
derived from opportunistically collected individuals, with unknown residence times.
133
Understanding the rate of radiocesium decline in affected aquatic systems is
imperative for elucidating potential risks to wildlife and humans who may consume these
vectors, as well as for assessing remediation efforts and determining when human
exclusion from contaminated areas is no longer necessary. Historical data for coots in the
SRS Pond B system provides a unique long-term data set for the calculation of the
decline of radiocesium in a natural system through the development of an ecological half-
life in biota. An ecological half-life delineates the decrease in contaminant concentration
over time within an ecosystem, and is thus reflected as a decrease in levels observed
within organisms as it becomes biologically unavailable (Brisbin 1993, Paller et al.
1999). The ecological half-life for radionuclides, like radiocesium, can be influenced by a
myriad of chemical, physical, and biological processes (Whicker and Schultz 1982,
Brisbin 1991a, Paller et al. 1999). Ecological half-life differs from the physical half-life,
which is the rate of radioactive decay to daughter products, and biological half-life,
which is the rate at which the isotope is turned over and eliminated from the body
(Brisbin 1993). A relatively long ecological half-life has previously been calculated for
Pond B biota (Table 4-1), possibly due to the low rate of water turnover in that system,
low potassium concentrations (thus radiocesium has less competition for binding in biota
and will remain available for accumulation instead of bound in the sediments), large
amounts of rooted aquatic macrophytes that can translocate cations from the sediment
into the water column, and the radiocesium remaining in the system is the 137Cs isotope
which has the 30.2 year half-life (Whicker et al. 1990).
As a variety of wildlife species can be exposed to radiocesium from Pond B,
investigating the ecological half-life is relevant not only for future access issues at SRS,
134
but also other locations dealing with the legacy of past nuclear pollution. Wetlands at
other nuclear weapons production sites (e.g. Willard 1960, Fitzner and Rickard 1975,
Halford et al. 1981) and in regions contaminated by nuclear accidents such as the
Chernobyl site (Brisbin 1991b) and Fukushima, Japan (Tagami and Uchida 2013) all
support populations of migratory waterfowl which can transfer radiocesium away from
these sites (Brisbin 1991a).
For this study, we collected American coots and ring-necked ducks from an
uncontaminated lake on SRS and released them on Pond B, then recollected birds
between approximately one and six months of radiocesium exposure. We included ring-
necked ducks because they can travel thousands of kilometers from the site (Kennamer
2003), and provide a comparative species with a more varied diet (Brisbin et al. 1973,
Bergan and Smith 1986, Hoppe et al. 1986). Our objectives were to 1) quantify uptake of
radiocesium at multiple time points of exposure by coots and ring-necked ducks
subsequent to translocation to Pond B and compare whole-body uptake between species
as well as concentrations in muscle and liver tissue, and 2) calculate the ecological half-
life for radiocesium in American coots on Pond B by comparing whole-body burdens to
historical data collected from the same location. These results have implications for
choosing focal species for experimental restriction to contaminated areas as well as for
long-term mensurative monitoring programs.
METHODS
Study Area
Research occurred on the Savannah River Site (SRS), an ~800 km2 limited-access former
nuclear production and research facility owned and operated by the U.S. Department of
135
Energy in South Carolina (White and Gaines 2000). The SRS was created in 1951 to
provide nuclear weapons materials (Savannah River Nuclear Solutions, LLC 2011), but
all five nuclear reactors have since been decommissioned (White and Gaines 2000).
Wetlands and other aquatic habitats on the SRS are abundant, including creeks, streams,
upland depressions, Carolina bays, bottomland and swamp forests, as well as two large
cooling reservoirs; these areas are important habitat for migrating waterfowl/waterbirds
(Lide 1994).
Pond B, a former recipient of heated effluent from R reactor on the SRS, is an 87-
ha reservoir with a mean depth of 4.3 m. Precipitation is the only input and the water has
low potassium levels and is slightly acidic (Figure 4-1; Alberts and Dickson 1985,
Alberts et al. 1988). Between 1961 and 1964, approximately 5.7 X 1012 Bq of
radiocesium was released into the Pond B reservoir (Ashley and Zeigler 1980). Use of
Pond B by 12 waterfowl species has been documented (Mayer et al. 1986). Radiocesium
in Pond B sediments (Brisbin et al. 1974) has been quantified and related to levels found
in migratory game birds with potential to disperse contaminants extensive distances from
the contaminated reservoir (Brisbin et al. 1973, Fendley et al. 1977, Brisbin and Vargo
1982, Kennamer et al. 1998). L-Lake, another cooling reservoir on the SRS, was used as
our uncontaminated source site for this research.
Trapping and Sample Collection
Many birds that reside or overwinter on the SRS are exposed to unknown
amounts of radiocesium contamination. Testing captured wild birds from contaminated
water sources does not evaluate residence time and birds can utilize multiple water
resources (contaminated and uncontaminated) during resident and migratory periods,
136
confounding estimates of contaminant uptake and exposure. To quantify accumulation of
radiocesium in coots and ring-neck ducks inhabiting Pond B, birds were trapped (coots:
December 7th 2013–January 13th 2014, coots and ring-necked ducks: December 8th–18th
2014) with welded-wire traps baited with corn (Haramis et al. 1982) or dip-netted from a
boat on L-Lake, an uncontaminated reservoir on the SRS. Captured birds were banded
with U.S. Fish and Wildlife Service aluminum leg bands, coots were fitted with colored
neck collars and ducks with colored nasal saddles for ease of identification during
collections, and a subset were whole-body counted for radiocesium to establish a
background level before release on Pond B. Before release we wing-scissored flight
feathers on one wing to render birds flightless, restricting individuals to the contaminated
waterbody.
Previous work in the Pond B system with coots found that they reach an
asymptotic radiocesium level (maximal radiocesium body burden) in approximately 20-
25 days (Brisbin et al. 1989, Brisbin 1991a). Therefore, birds were left on Pond B a
minimum of 30 days before lethal collection (using shotguns) to ensure maximum body
burdens would be reached in coots for comparison with previous data. Collections were
conducted at approximately 30, 60, and 90 days (with a few birds left for later recovery
up to 173 days) to verify achievement of asymptotic body burdens in coots and ring-
necked ducks at approximately 30 days of exposure. Upon collection birds were weighed
and frozen at -20°C for later whole-body counting and dissection. All birds were whole-
body counted for radiocesium and then dissected to collect breast muscle and liver tissues
for radiocesium quantification in specific tissues. Wet weights of all muscle and liver
samples were recorded before samples were freeze-dried, then weighed again before
137
being homogenized into a powder using a coffee grinder. Grinder canisters were cleaned
(5% nitric acid solution) and dried between uses. All animal handling practices and
euthanasia were carried out with accompanying federal and state collecting permits and
in accordance with University of Georgia Animal Care and Use guidelines under protocol
A2013 06-004-Y3-A1.
Radiocesium Analysis
Birds were whole-body counted with a 10·2-cm x 15·2-cm NaI (Tl) gamma
detector (Bicron Model:6H3Q/5; S/N:BJ-124R) coupled to an IBM 300-GL Personal
Computer (Windows 98 OS) containing an onboard Canberra MCA card and controlled
by Canberra Genie 2000 gamma spectroscopy software (Version 1.3; entire system
located in SREL Lab 120). A counting window (Region of Interest-ROI) of 596–728
kiloelectron-Volts (keV) centered on 662 keV was used to record total detector
absorption events from the radiocesium emission of 662 keV photons. The system was
calibrated daily, as counting took place, with a traceable radiocesium calibration disc
(New England Nuclear Gamma Reference Disc Source Set; Catalogue No. NES-101S;
radiocesium disc; 1.04 microCuries on 10/2/1985) by adjusting the system amplifier gain
control to center the disc-generated peak on channel 331 (661.7keV). Generally, 30-min
count times (1800 sec) were used for counting collected/sacrificed birds (whole-body;
frozen) and backgrounds (empty chamber), while 15-min count times (900 sec) were
used for counting aqueous standards (the ILB Series of standards; containing known
radiocesium quantities [Becquerels (Bq)]; decay corrected to count dates). Background-
corrected count rates (counts per second; cps) from the ILB Series of standards were used
to produce mass-specific count yields which were in turn used to produce a predictive
138
equation of expected yields from bird mass (in grams; yield=0.4449*mass-0.343;
R2=0.97). Finally, background-adjusted bird count rates and the birds' mass-specific
yields were used to determine radiocesium content of birds (whole-body Bq, Bq/g,
pCi/g).
All muscle and liver samples were freeze-dried, powdered, then packed into
scintillation vials for quantification of radiocesium activity using a Packard Cobra II
Auto-Gamma Counter (Model Cobra II 5003) with a single 3-inch through-hole NaI
detector. A counting window (Region of Interest-ROI) of 580–754 kiloelectron-Volts
(keV) centered approximately on 662 keV was used to record total absorption events
from the radiocesium emission of 662 keV photons. The system was auto-calibrated
daily, as counting took place, with a traceable radiocesium source (SREL-0113; 0.1
microCuries on 10/2012). Generally, 60-min count times (3600 sec) were used for
counting dry, powdered samples (packed into tubes) and backgrounds (empty tubes) that
were arranged in every fifth counting position. Four standards were prepared from
commercially-available chicken breast muscle tissue that was dried, homogenized into
powder, spiked with known quantities of radiocesium (745 Bq in each spiked standard;
decay-corrected for the date of preparation), and then loaded/packed into 4 scintillation
tubes in 1-gram increments ranging from 1-4g. Background-corrected count rates (net
counts per second; ncps) recorded for these spiked standards were used to produce
estimates of mass-specific count yields (a ratio of measured net count rate [ncps] to
expected disintegration rates [dps] for the known radiocesium activity) at each of 4
available sample height settings relative to the NaI detector of the Cobra II system.
Sample position #4 produced count yields that varied little across all sample masses (SD
139
< 0.007) and so samples/backgrounds were all counted using sample height #4 setting,
with an averaged count yield value of 0.2213 used as a constant in radiocesium content
determinations for unknowns. Specifically, background-adjusted dry tissue count rates
divided by the yield constant were estimated as the radiocesium content of dry tissue
samples (total Bq, Bq/g [dry mass], and Bq/g [wet mass]). Minimum Detectable
Concentrations (MDCs) were calculated for all radiocesium analyses using the equation
of Lloyd Currie (Currie 1968).
Data Analysis
Many of the whole-body counts of radiocesium for the birds that were captured
on L-Lake were below background (or blank count) rates, which gave us negative values
for radiocesium concentrations. A common practice is to not report these values, or the
small positive values that happen to be below MDCs, or set them to an arbitrary indicator
value, thus biasing the mean and variance of the concentration data (Gilbert and Kinnison
1981, Newman et al. 1989). We chose to report all values, including negatives and those
below the MDCs, when determining average radiocesium concentrations and standard
errors (SE).
All statistical tests were performed with JMP Pro Version 10.0.1; SAS Institute
Inc., Cary, NC. We used t-tests to compare whole-body, muscle, and liver radiocesium
concentrations between coots and ring-necked ducks, after testing for equality in the
variances. To evaluate the relationship between whole-body radiocesium concentrations
and concentrations in muscle and liver, we used linear regression (with forced intercepts
of zero). We compared the slope estimates of the relationships between whole-body and
tissue burdens for each species with homogeneity of slopes models (Analysis of Variance
140
models using least squares fits). Using the common muscle slope for both species we
calculated the whole-body radiocesium equivalent, having this standard for the muscle to
whole-body relationship allows a comparison to the EEC limit for radiocesium in fresh
meat (0.600 Bq/g), with only having to whole body count birds, which is much less labor
intensive than completing muscle counts.
From historical coot collection data on Pond B during winters of 1975-76, 1983-
84, 1986-87, 1987-88, and 1991-92 along with our coot data collected from Pond B in
winters 2013-2014 and 2014-2015 (Figure 4-2a), we modeled the long-term natural
attenuation of radiocesium body burdens (Brisbin and Kennamer 2000). The last input of
radiocesium was in 1965, and we would expect a decline of radiocesium through time
since then; however investigation of historical data revealed that radiocesium burdens in
coots increased between the winters of 1975-76 and 1986-87, possibly from residual
radiocesium flushing out of the canal that leads from the reactor to the reservoir during
flooding events. Therefore, we chose to model the ecological half-life in the system from
the (natural log-transformed) whole-body radiocesium data for 96 coots starting at the
winter of 1986-87 through our present day collection data, in winters 2013-2014 and
2014-2015 (Figure 4-2b). Using linear regression we estimated model parameters and
95% confidence intervals (CI) to describe the decline of radiocesium whole-body burdens
(Bq/g wet mass; y) for Pond B coots sampled during different years (x; continuous
variable). Our estimated parameters included the slope (β) and the intercept (ea , e is the
base of the natural log). Since 1965 was the initial winter year migratory coots would
have been exposed to the maximum radiocesium concentration on Pond B, we set the
intercept to this year (Ashley and Zeigler 1980). The slope acquired from our regression
141
model was used in the half-life equation for radiocesium (T1/2 =ln(2) / β) to produce an
ecological half-life (Te) estimate and a 95% CI around the estimate.
RESULTS
Upon initial capture from L-Lake in 2013-2015 a random selection of 32 birds (25
coots and 7 ring-necked ducks) were whole-body counted to establish a background
concentration of radiocesium before moving them to Pond B. These birds had average
radiocesium concentrations of 0.0006 Bq/g (coots) and -0.0014 Bq/g (ring-necked
ducks), with ranges of -0.0022–0.0051 Bq/g and -0.0046–0.0001 Bq/g, respectively
(Table 4-2). All whole-body concentrations were below individual corresponding MDCs
of 0.012 Bq/g for coots and 0.011Bq/g fresh mass for ring-necked ducks (Table 4-2).
In the winter of 2013–2014 we released 34 coots to Pond B and re-collected 16
between 44 to 88 days of exposure. During winter of 2014-2015 we released 50 coots and
55 ring-necked ducks on Pond B and re-collected birds between 32 to 173 days of
exposure. In total, we were able to recollect 35 coots and 36 ring-necked ducks, whose
whole-body radiocesium concentrations averaged 0.94±0.064 SE Bq/g and 1.18±0.097
SE Bq/g, respectively (Table 4-3). All collected birds had whole-body radiocesium
concentrations above their respective MDCs (0.014 Bq/g for coots and 0.012 Bq/g fresh
mass for ring-necked ducks; Table 4-3). A t-test (unequal variances), showed that whole-
body radiocesium concentrations differed significantly between species (t = 2.04, P =
0.05), with ring-necked ducks having higher concentrations.
Congruent with previous studies of coot attenuation of radiocesium on Pond B, in
our study coots maintained equilibrium whole-body concentrations from the first
collection event at 32 days through 90 days of exposure (Figure 4-3a). A small sample
142
size of coots (N=5) collected beyond 90 days of exposure suggested eventual decline in
whole-body radiocesium concentrations (Figure 4-3a). Contrary to the equilibrium
reached at approximately 30 days in coots, whole-body radiocesium concentrations in
ring-necked ducks continued to increase between 33 and 90 days of exposure, before we
saw eventual decline beyond 90 days (Figure 4-3b).
Muscle and liver radiocesium concentrations for the 35 coots averaged
1.64±0.118 SE Bq/g and 1.02±0.062 SE Bq/g wet mass, respectively (Table 4-3). In the
36 ring-necked ducks muscle and liver concentrations averaged 2.24±0.195 SE Bq/g and
1.25±0.097 SE Bq/g wet mass, respectively (Table 4-3). Ring-necked ducks had higher
concentrations of radiocesium levels in muscle tissue (t = 2.64, P = 0.01), but liver levels
were not different between species (t = 1.94, P = 0.06). All coots and ring-necked ducks
had muscle and liver radiocesium levels above their respective MDCs (Table 4-3).
Ninety-seven percent (34 of 35) of coots and 92% (33 of 36) ring-necked ducks had
muscle concentrations above the European Economic Community (EEC 1986) limit of
0.600 Bq/g in fresh meat established for human consumption (Figure 4-3).
In coots and ring-necked ducks radiocesium concentrations in wet muscle and
liver tissues were positively correlated to whole-body radiocesium burdens (Figure 4-4).
Results indicated that radiocesium concentrated to a greater extent into muscle (β =
1.74±0.036 SE, β = 1.91±0.030 SE) than into liver (β = 1.06±0.028 SE, β = 1.03±0.034
SE) for both coots and ring-necked ducks, respectively. In comparing slopes for the
relationships of tissue concentrations to whole-body radiocesium between species we
found no significant relationship for muscle (overall model: F3,67 = 441.1, P < 0.0001, R2
= 0.95; slopes: F1,67 = 0.002, P = 0.10) or liver (overall model: F3,67 = 101.5, P <0.0001,
143
R2 = 0.82; slopes: F1,67 = 0.002, P = 0.97). Because the slopes were not different we
developed a common slope for both species combined (muscle: β = 1.85±0.025 SE, liver:
β = 1.04±0.023 SE). We estimated the whole-body radiocesium equivalent for both
species using the common muscle slope (β = 1.85), suggesting a whole-body burden of
0.324 Bq/g, would be necessary to exceed the EEC limit (0.600 Bq/g) for radiocesium in
fresh muscle. Using linear regression (R2 = 0.31) we obtained an α estimate of 1.84±0.25
SE which allowed us to estimate the mean radiocesium concentration in Pond B coots
after the releases into the cooling water in 1965 at 6.30 Bq/g (95% CI = 3.86-10.28). The
β estimate (-0.0412±0.0064 SE) when incorporated into the half-life equation (T1/2 =ln(2)
/ β) produced an ecological half-life (Te) estimation of 16.82 (95% CI = 12.91-24.19)
years.
DISCUSSION
We verified historical research reporting that whole-body radiocesium burdens in
coots reached equilibrium with the environment in ≤30 days of exposure (Brisbin et al.
1989), although our data suggest burdens may have begun to decline after about 90 days
of exposure. In contrast, whole-body radiocesium concentrations in ring-necked ducks
continued to increase up to approximately 90 days before also experiencing a gradual
decline. Past research with free-ranging coots observed a decline in radiocesium burdens
in late spring, speculating this trend may have been due to the “dilution” of those birds
that had spent the winter on the contaminated water body (but were leaving north on
migration) with migrating birds returning north from uncontaminated waters during
migration (Brisbin et al. 1991a). However, given that we observed a similar decline
towards the end of our experiment (spring) in both species of restricted birds, we
144
hypothesize this decline may be due to excretion through a seasonal change in physiology
or bioavailability through altered dietary habits. Additional research into temporal
patterns of radiocesium uptake and elimination is therefore needed to more clearly
elucidate potential seasonal and dietary effects on radiocesium accumulation to improve
risk assessment calculations of free-ranging birds.
Our findings revealed ring-necked ducks had higher concentrations of
radiocesium in muscle tissue than coots, possibly due to differences in diet or other
behaviors that may influence exposure, but liver concentrations did not differ between
species. Muscle, where radiocesium accumulates to a higher degree then organs (Potter et
al. 1989), is used when determining human consumption risks for contaminated meat.
When comparing muscle concentrations from both species to the EEC recommended
limits for consumption of radiocesium in fresh meat, we found that 97% of coots and
92% of ring-necked ducks in our study were above these levels. In a previous study with
free-ranging birds opportunistically sampled from Pond B, only 63.3% were above this
threshold and coots from Pond B had whole-body radiocesium levels about 7 times
higher than diving ducks [ruddy ducks (Oxyura jamaicensis) and bufflehead (Bucephala
albeola)] sampled at the same time (Chapter 2). These data suggest that by restricting
birds we increased residency time (and thus radiocesium accumulation) beyond what
would normally be expected in free-ranging populations. These data also suggest that
coots may spend more time foraging in this habitat than diving ducks, although ring-
necked ducks were not included in the previous assessment of free-ranging birds on Pond
B (Chapter 2). Given the variation observed in uptake and accumulation between species
and discrepancies with previous mensurative studies, our data suggest a combination of
145
mensurative and experimental research with more than one species are needed in
ecological monitoring of radiocesium.
A previous study used muscle concentrations of a small sample of Pond B coots
to estimate the whole-body equivalent concentration of radiocesium (0.47 Bq/g), this
value is the amount of radiocesium which if counted in the whole-body of subsequently
collected coots then equates to the EEC suggested limit for fresh meat of 0.600 Bq/g
(Potter 1987). However, the whole-body radiocesium equivalent calculated from the
combined muscle slope for both species in our study was 0.324 Bg/g. This shows that the
updated whole-body radiocesium concentration (lower in radiocesium than the past
calculation) equates to the EEC limit.
With the physical half-life of 137Cs being approximately 30 years, an estimate for
how long the isotope would be elevated above background levels in the system is
generally 5 times as long as the physical half-life (in this case ~150 years after release
before 137Cs would no longer be detectable because of normal radioactive decay
processes; Brisbin 1991a). However, because of later release of radiocesium bound in
sediments or vegetation or its disappearance from the local environment through
migration, weathering, and removal by organisms, the ecological half-life can differ from
the physical half-life. In the SRS Pond B system, our estimated ecological half-life
suggest it will be approximately 17 years (possible range between 13-24 years) before the
whole-body equivalent for radiocesium reaches a level that is under the EEC limit for
fresh meat. Thus, if the ecosystem remains stable (i.e. the bioavailability of 137Cs does not
change), and residence times remain similar for birds, it will not be until approximately
146
2030 before whole-body radiocesium fall below possible restrictive limits for human
consumption.
CONCLUSION
The risks associated with radionuclides are of ecological concern because of the
long physical half-lives of certain isotopes like radiocesium (137Cs), remaining as
environmental contaminants available for uptake by biota (Hinton 1998). Given that
radiocesium concentrates in skeletal muscle and thus may be consumed by hunters,
understanding rates of uptake and elimination of radionuclides requires basic and applied
research, through mensurative and experimental exposure studies in natural conditions, to
fully comprehend potential effects on wildlife and risks to human consumers. Research in
areas with a long history of contamination can serve as a model for areas just beginning
long-term ecological risk assessments, such as Fukushima, Japan and long-term data used
to derive ecological half-life calculations can be useful in predicting the length of time
and degree to which contaminated systems pose health or environmental risks. Additional
studies are needed to elucidate the effects of residence time on radiocesium accumulation
for a greater diversity of waterfowl/waterbirds and other wildlife known to utilize
radionuclide-contaminated sites worldwide. Collectively, the data presented herein
suggest that waterfowl/waterbirds represent a possible vector for human exposure to
radionuclides from contaminated areas, especially areas in which birds may stop-over on
migration or areas in which they over-winter for months at a time.
147
Table 4-1. Ecological half-life estimates for species from Pond B or Par Ponda on the
Savannah River Site (SRS). (For the current study American coots were restricted to
Pond B for between 33 and 173 days of exposure to radiocesium in that system.)
Species Location Ecological Half-Life
(yrs)
95% CI (yrs) Citation
American coot Pond B 16.8 12.9-24.2 Current study Largemouth bass Pond B 16.7 14.3-20.3 Paller et al. 1999, 2002
Largemouth bass Pond B 13.6 N/A Peles et al. 2000
Sunfishes Pond B 13.4 8.1-47.2 Paller et al. 1999, 2002
American coot Par Pond-North Arm 4.92 4.4-5.6 Brisbin and Kennamer 2000
American coot Par Pond-Hot Arm 3.87 3.5-4.4 Brisbin and Kennamer 2000
American coot Par Pond-West Arm 4.36 3.7-5.3 Brisbin and Kennamer 2000
Largemouth bass Par Pond 4.99 3.7-7.9 Paller et al. 1999, 2002
Sunfishes Par Pond 4.78 3.5-7.4 Paller et al. 1999, 2002 a All ecological half-life estimates for Par Pond were from periods before the 1991 drawdown of
the reservoir.
148
Table 4-2. Descriptive statistics for a random sampling of American coots and ring-necked ducks that were trapped from L-
Lake and whole-body counted for radiocesium prior to release onto Pond B on the Savannah River Site (SRS) over the winter
of 2013-2015.
Species Variable n % below MDC a Mean SE Min Max Median CV (%) American coot Bq/g; fresh mass 25 100 0.0006 0.0004 -0.0022 0.0051 0.0002 294.6
MDC; fresh mass 25 100 0.012 0.0003 0.011 0.016 0.012 11.2
Ring-necked duck Bq/g; fresh mass 7 100 -0.0014 0.0007 -0.0046 0.0001 -0.0007 -123.3 MDC; fresh mass 7 100 0.011 0.0002 0.01 0.012 0.011 4.2
a MDC=minimum detectable concentration; calculated from Currie (1968)
149
Table 4-3. Descriptive statistics for radiocesium concentrations of American coots and ring-necked ducks that were released to
Pond B on the Savannah River Site (SRS) for between 33 and 173 days of exposure before being collected.
Species Sample Variable n % below MDC a Mean SE Min Max Median CV (%)
American coot Whole-body Bq/g; fresh mass 35 0 0.941 0.064 0.19 1.85 0.885 39.9
MDC; fresh mass 35 0 0.014 0.0002 0.012 0.017 0.014 10.5
Muscle Bq/g; dry mass 35 0 6.258 0.448 1.248 11.805 5.95 42.4
Bq/g; wet mass 35 0 1.636 0.118 0.279 3.006 1.559 42.7
MDC; dry mass 35 0 0.352 0.0057 0.303 0.436 0.353 9.5
wet:dry ratio 35 0 3.814 0.043 3.423 4.53 3.756 6.7
Liver Bq/g; dry mass 35 0 3.554 0.213 0.668 5.711 3.484 35.4
Bq/g; wet mass 35 0 1.024 0.062 0.192 1.69 1.014 36
MDC; dry mass 35 0 0.318 0.0052 0.273 0.388 0.31 9.7
wet:dry ratio 35 0 3.478 0.03 3.16 3.874 3.448 5.2
Ring-necked duck Whole-body Bq/g; fresh mass 36 0 1.178 0.097 0.122 2.18 1.226 49.4
MDC; fresh mass 36 0 0.012 0.0001 0.011 0.014 0.012 4.2
Muscle Bq/g; dry mass 36 0 7.495 0.661 0.944 15.447 7.339 52.9
Bq/g; wet mass 36 0 2.238 0.195 0.28 4.55 2.282 52.4
MDC; dry mass 36 0 0.321 0.0043 0.273 0.369 0.314 8.1
wet:dry ratio 36 0 3.347 0.018 3.157 3.691 3.323 3.3
Liver Bq/g; dry mass 36 0 4.311 0.331 0.707 7.583 4.159 46
Bq/g; wet mass 36 0 1.247 0.097 0.184 2.235 1.211 46.5
MDC; dry mass 36 0 0.317 0.0048 0.257 0.368 0.316 9
wet:dry ratio 36 0 3.455 0.031 3.137 3.987 3.439 5.5 a MDC=minimum detectable concentration; calculated from Currie (1968)
150
Figure 4-1. The Par Pond Reservoir system on the Savannah River Site (SRS) that
includes P- and R-reactors with depictions of canals that carried the radionuclide
contaminated cooling water to Ponds B and C and Par Pond during several reactor
releases.
151
a.)
b.)
Figure 4-2. Historical and current data for whole-body radiocesium concentrations in
American coots from Pond B on the Savannah River Site between the winters of 1975-
1976 and 2014-2015; a.) dashed line for the whole-body equivalent (0.324 Bq/g) to the
European Economic Community limit for radiocesium in fresh meat (0.600 Bq/g) and b)
a linear regression of natural log-transformed data from collections between winters
1986-1987 and 2014-2015, estimates utilized in ecological half-life calculations.
0.0
2.0
4.0
6.0
8.0
10.0
12.0 W
hole
-bod
y 137C
s (B
q/g
fres
h m
ass)
Winter of Collections
EEC fresh meat 137
Cs limit (whole-body equivalent)
-3.0
-2.0
-1.0
0.0
1.0
2.0
3.0
Log
Who
le-b
ody
137 C
s (B
q/g
fres
h m
ass)
Winter*of*Collections
α estimate: 1.84 ± 0.25 SE β estimate: -0.0412 ± 0.0064 SE R
2 = 0.31
152
a.)
b.)
Figure 4-3. Whole-body radiocesium concentrations in a.) American coots and b.) ring-
necked ducks collected from Pond B on the Savannah River Site (SRS) after exposure
between 32 and 173 days. Day 0 whole-body concentrations are counts done on live-
captured birds from L-Lake before release onto Pond B. Solid lines show non-linear fits
to the data and dashed lines represent our calculated whole-body equivalent (0.324 Bq/g)
to the European Economic Community limit for radiocesium in fresh meat (0.600 Bq/g).
0.0
0.5
1.0
1.5
2.0
2.5
0 30 60 90 120 150 180
Who
le-b
ody
137 C
s (B
q/g
fres
h m
ass)
Days post-release
EEC fresh meat 137
Cs limit (whole-body equivalent)
0.0
0.5
1.0
1.5
2.0
2.5
0 30 60 90 120 150 180
Who
le-b
ody
137 C
s (B
q/g
fres
h m
ass)
Days post-release
EEC fresh meat 137
Cs limit (whole-body equivalent)
153
a.)
b.)
Figure 4-4. Whole-body radiocesium concentrations in a.) American coots and b.) ring-
necked ducks collected from Pond B on the Savannah River Site (SRS) after exposure
between 32 and 173 days. Day 0 whole-body concentrations are counts done on live-
captured birds from L-Lake before release onto Pond B. Solid lines show non-linear fits
to the data and dashed lines represent our calculated whole-body equivalent (0.324 Bq/g)
to the European Economic Community limit for radiocesium in fresh meat (0.600 Bq/g).
0.0
0.5
1.0
1.5
2.0
2.5
0 30 60 90 120 150 180
Who
le-b
ody
137 C
s (B
q/g
fres
h m
ass)
Days post-release
EEC fresh meat 137
Cs limit (whole-body equivalent)
0.0
0.5
1.0
1.5
2.0
2.5
0 30 60 90 120 150 180
Who
le-b
ody
137 C
s (B
q/g
fres
h m
ass)
Days post-release
EEC fresh meat 137
Cs limit (whole-body equivalent)
154
LITERATURE CITED
Alberts, J. J., J. W. Bowling, J. E. Schindler, and D. E. Kyle. 1988. Seasonal dynamics of
physical and chemical properties of a warm monomictic reservoir. Internationale
Vereinigung fuer Theoretische und Angewandte Limnologie 23:176-180.
Alberts, J. J., and T. J. Dickson. 1985. Organic carbon and cation associations in humic
material from pond water and sediment. Organic Geochemistry 8:55-64.
Ashley, C., and C. Zeigler. 1980. Releases of radioactivity at the Savannah River Plant,
1954 through 1978. Report DPSU 75-25-1. E.I. DuPont deNemours, Savannah
River Laboratory, SC.
Bergan, J. F., and L. M. Smith. 1986. Food robbery of wintering Ring-necked Ducks by
American Coots. The Wilson Bulletin 98:306-308.
Bagshaw C., I. L. Brisbin Jr. 1984. Long-term declines in radiocesium of two sympatric
snake populations. Journal of Applied Ecology 21:407-413.
Brisbin Jr., I. L., R. A. Geiger, and M. H. Smith. 1973. Accumulation and redistribution
of radiocesium by migratory waterfowl inhabiting a reactor cooling reservoir.
Environmental behavior of radionuclides released in the nuclear industry,
International Atomic Energy Agency symposium (IAEA-SM-172/72) Vienna,
Austria, pp. 373-384.
Brisbin Jr., I. L., R. J. Beyers, R. W. Dapson, R. A. Geiger, J. B. Gentry, J. W. Gibbons,
M. H. Smith, and S. K. Woods. 1974. Patterns of radiocesium in the sediments of
a stream channel contaminated by production reactor effluents. Health Physics
27:19-27.
155
Brisbin Jr., I. L., and M. J. Vargo. 1982. Four-year declines in radiocesium
concentrations of American coots inhabiting a nuclear reactor cooling reservoir.
Health Physics 43:266-269.
Brisbin Jr., I. L., M. C. Newman, S. G. McDowell, and E. L. Peters. 1989. The prediction
of contaminant accumulation by free-living organisms: Applications of a
sigmoidal model. Environmental Chemistry and Toxicology 9:141-149.
Brisbin Jr., I. L. 1991a. Avian radioecology. Current Ornithology, Volume 8. (Ed. D.M.
Power) Plenum Publishing Corporation, New York, NY, pp. 69-140.
Brisbin Jr., I. L. 1991b. Birds as indicators of global contamination processes: The
Chernobyl connection. In Acta XX Congressus Internationalis Ornithologici, New
Zealand Ornithological Congress. Wellington, NZ, pp. 2503-2508.
Brisbin Jr., I L. 1993. Birds as monitors of radionuclide contamination. Birds as monitors
of environmental change. (Ed. R. W. Furness) Chapman and Hall, London, UK.
Brisbin Jr., I. L., and R. A. Kennamer. 2000. Long-term studies of radionuclide
contamination of migratory waterfowl at the Savannah River Site: Implications
for habitat management and nuclear waste site remediation. Studies in Avian
Biology 21:57-64.
Brisbin Jr., I. L., H. D. Pratt, and T. B. Mowbray. 2002. American coot (Fulica
americana) and Hawaiian coot (Fulica alai). The Birds of North America No.697.
(Eds. A. Poole and F. Gill) The Academy of Natural Sciences, Philadelphia,
Pennsylvania.
156
Carlton, W. H., L. R. Bauer, A. G. Evans, L. A. Geary, C. E. Murphy Jr., J. E. Pinder,
and R. N. Strom. 1992. Cesium in the Savannah River Site environment. No. DE-
AC09-89SR18035 Westinghouse Savannah River Company, Aiken, SC.
Carlton, W. H., C. E. Murphy Jr., and A. G. Evans. 1994. Radiocesium in the Savannah
River Site Environment. Health Physics 67:233-244.
Cristol, D. A., L. Savoy, D. C. Evers, C. Perkins, R. Taylor, and C. W. Varian-Ramos.
2012. Mercury in waterfowl from a contaminated river in Virginia. Journal of
Wildlife Management 76:1617-1624.
Currie, L. A. 1968. Limits for qualitative detection and quantitative determination.
Analytical Chemistry 40:586-593.
European Economic Community (EEC). 1986. Derived reference levels as a basis for the
control of food stuffs following a nuclear accident: A recommendation from the
group of experts set up under Article 31 of the Eurotom Treaty. EEC Regulation
1701/86, Commission of the EEC, Printing Office, Brussels, Belgium.
Evans, D.W., J. J. Alberts, and R. A. Clark. 1983. Reversible ion-exchange fixation of
cesium-137 leading to mobilization from reservoir sediments. Geochimica et
Cosmochimica Acta 47:1041–1049.
Fendley, T. T., M. N. Manlove, and I. L. Brisbin Jr. 1977. The accumulation and
elimination of radiocesium by naturally contaminated wood ducks. Health Physics
32: 415-422.
Fitzner, R. E., and W. H. Rickard. 1975. Avifauna of waste ponds ERDA Hanford
Reservation, Benton County, Washington. Battelle Pacific Northwest Labs
Report.
157
Gilbert, R. O., and R. R. Kinnison. 1981. Statistical methods for estimating the mean and
variance from radionuclide data sets containing negative, unreported or less-than
values. Health Physics 40:377-390.
Halford, D. K., J. B. Millard, and D. O. Markham. 1981. Radionuclide concentrations in
waterfowl using a liquid radioactive waste disposal area and the potential
radiation dose to man. Health Physics 40:173-181.
Haramis, G. M., E. L. Derleth, and D. G. McAuley. 1982. Techniques for trapping, aging,
and banding wintering canvasbacks. Journal of Field Ornithology 53:342-351.
Hinton, T. G. 1998. Estimating human and ecological risks from exposure to radiation.
Risk Assessment: Logic and Measurement (Eds. M. C. Newman, and C. L.
Strojan) Ann Arbor Press, Chelsea, MI, pp. 143-166.
Hoppe, R. T., L. M. Smith, and D. B. Wester. 1986. Foods of wintering diving ducks in
South Carolina. Journal of Field Ornithology 57:126-134.
Kennamer, R. A., I. L. Brisbin Jr., C. D. McCreedy, and J. Burger. 1998. Radiocesium in
Mourning Doves: Effects of a contaminated reservoir drawdown and risk to
human consumers. Journal of Wildlife Management. 62:497-508.
Kennamer, R. A. 2003. Recoveries of ring-necked ducks banded on the U.S. Department
of Energy’s Savannah River Site, South Carolina. The Oriole 68:8-14.
Kennamer, R. A. 2005. Waterfowl. Ecology and management of a forested landscape:
Fifty years on the Savannah River Site. (Eds. J. C. Kilgo and J. I. Blake) Island
Press, Washington, DC, pp. 347-359.
158
Lide, R. F. 1994. Carolina bays and similar natural depression wetlands of the SRS. GIS
coverage in ARC/INFO. Savannah River Ecology Laboratory, Savannah River
Site, Aiken, SC.
Mayer, J. J., R. A. Kennamer, and R. T. Hoppe. 1986. Waterfowl of the Savannah River
Plant. Final Report. Report SREL-22 UC-66e. Savannah River Ecology
Laboratory, Division of Stress and Wildlife Ecology, Aiken, SC.
Mohler, H. J., F. W. Whicker, and T. G. Hinton. 1997. Temporal trends of 137Cs in an
abandoned reactor cooling reservoir. Journal of Environmental Radioactivity
37:251-268.
Newman, M. C., P. M. Dixon, B. B. Looney, and J. E. Pinder, III. 1989. Estimating mean
and variance for environmental samples with below detection limit observations.
Water Resources Bulletin 25:905-916.
Paller, M. H., J. W. Littrell, and E. L. Peters. 1999. Ecological half-lives of 137Cs in fishes
from the Savannah River Site. Health Physics 77:392-407.
Peles, J. D., A. L. Bryan Jr., C. T. Garten Jr., D. O. Ribble, and M. H. Smith. 2000.
Ecological half-life of 137Cs in fish from a stream contaminated by nuclear reactor
effluents. Science of the Total Environment 263:255-262.
Potter, C. M. 1987. Use of reactor cooling resevoirs and cesium-137 uptake in the
American coot. M. S. thesis, Colorado State University, Fort Collins, CO.
Potter, C. M., Brisbin, I. L., Jr., McDowell, S. G., Whicker, F. W., 1989. Distribution of
137Cs in the American coot (Fulica americana). J. Environmental Radioactivity 9,
105-115.
Savannah River Nuclear Solutions, LLC. 2011. Environmental, safety, health, and quality
159
regulatory integration and environmental services. Environmental Management
System (EMS) description manual.
http://www.srs.gov/general/pubs/envbul/documents/ems_manual.pdf Accessed
15 October 2014.
Tagami, K., and S. Uchida. 2013. Radiocesium concentration change in game animals:
Use of food monitoring data. Waste Management Conference: International
collaboration and continuous improvement INIS-US--13-WM-13168.
Whicker, F. W., and V. Schultz. 1982. Radioecology: Nuclear energy and the
environment (2 vols). CRC Press, Boca Raton, FL.
Whicker, F. W., J. E. Pinder III, J. W. Bowling, J. J. Alberts, and I. L. Brisbin Jr. 1990.
Distribution of long-lived radionuclides in an abandoned reactor cooling
reservoir. Ecological Monographs 60:471-496.
White, D. L., and K. E. Gaines. 2000. The Savannah River Site: Site description, land
use, and management history. Studies in Avian Biology 21:8-17.
Willard, W. K. 1960. Avian uptake of fission products from an area contaminated by
low-level atomic wastes. Science 132:148-150.
160
CHAPTER 5
CONCLUSION
To expand our knowledge of the levels of trace elements and radiocesium that
game animals accumulate from contaminated areas on the Savannah River Site (SRS),
and similarly contaminated sites worldwide, the overarching objectives of my research
were to: 1) quantify levels of contaminants in common game species from areas with a
range of levels of contaminants, 2) investigate trace element/radiocesium accumulation
rates in waterfowl/water birds inhabiting contaminated habitats, and 3) relate mensurative
and experimental contaminant burdens to levels potentially harmful to wildlife and to
derive human consumption limits for several taxa inhabiting contaminated ecosystems.
To accomplish this research I studied contaminant burdens in both lethal (muscle, liver)
and non-lethal tissues (blood, whole-body burdens) for several species with the additional
goal of developing predictive relationships to facilitate non-destructive sampling in future
studies.
In Chapter 2, I conducted a mensurative study to quantify levels of trace elements
and radiocesium in wild pigs (Sus scrofa), gray squirrels (Sciurus carolinensis), and a
variety of waterfowl/waterbird species on the SRS. Contaminant burdens for SRS wild
pigs were further compared to wild pigs collected from a non-contaminated area of
central Georgia. I found SRS pigs had significantly higher burdens of Hg, Se, and Zn,
while Georgia pigs had significantly higher Cr. SRS pigs also had significantly higher
levels of radiocesium in liver and muscle, with muscle burdens being 5 times higher in
161
SRS pigs. Evaluation of contaminant burdens in squirrels revealed there were no
differences in trace element levels among individuals collected across my four sampling
sites on the SRS, but radiocesium concentrations were significantly higher in areas with
known radiocesium contamination (Pond B and Fourmile branch beaver pond).
For diving ducks, I compared trace element tissue concentrations between D-Area
(CCW settling basin site), and all other SRS locations and found that D-Area waterfowl
had significantly higher Se (muscle and liver) and Zn (muscle) burdens, while diving
ducks from the other locations had higher Cu in liver and Cr in muscle. Radiocesium
concentrations from birds collected from Pond B, a reservoir contaminated with
radionuclides, were significantly higher than for birds collected at non-radiologically
contaminated sites (Fourmile, L-Lake, and D-Area; no differences existed between these
locations). Human consumption limits based on As, Se, and Hg burdens in muscle and
Hg EPA dose limits showed that for wild pigs and squirrels the number of allowed meals
exceeded numbers that hunters in the region generally consume. However, diving ducks
sampled from the CCW settling basins had much lower meal allowances; at average trace
element concentrations in muscle adults would be allowed 2.1 meals per month and
children 1 meal per month; while at maximum concentrations 1/3 of a meal and 1/10 of a
meal would be allowed for adults and children, respectively. No wild pigs or squirrels
exceeded the European Economic Community limit for radiocesium in fresh meat, but 19
of the 98 tested waterfowl/waterbirds did exceed the limit, all of whom were collected
from Pond B.
In Chapter 3, I conducted an experiment where I trapped ring-necked ducks
(Aythya collaris) from an uncontaminated lake and after banding, affixing colored nasal
162
saddles, and clipping their flight feathers on one wing released them into the D-Area ash
basins. I subsequently re-collected these birds between 3 to 92 days of exposure to
elucidate accumulation rates of several contaminants of concern found in CCW. I
observed a positive linear trend through time in the accumulation of Se in muscle, liver,
and blood, with the strongest relationship in muscle (R2 0.76). Hg had a significant
positive relationship with days of exposure in muscle (R2 0.14) and a significant negative
relationship with days of exposure in liver (R2 0.19). Levels of Se accumulation caused
human consumption limits to drop from 26.7 to 6.2 meals per month from the first
exposure period (3-15 days exposure) to the last (76-92 days exposure) for adults and
drop from 12.2 to 2.8 meals per month for children at average muscle concentrations. Se
in liver was correlated with As and Hg, while no correlations among the elements were
found in muscle. Se and As levels in muscle were correlated to levels in observed in liver.
Also, I was able to relate blood sample concentrations of As to concentrations in muscle
and liver, as well as Se in blood to muscle, and create equations that would allow
prediction of muscle and liver concentrations from blood contaminant concentrations.
These data suggested a non-lethal blood sample, which may be able to be added to
existing USFWS banding protocols to provide a country-wide dataset for monitoring
trace element accumulation and potential risk of exposure to hunters across the 3 main
flyways, may be able to be used to test for As in muscle and liver and Se in liver with
reasonable predictive power. It is possible a predictive equation could be created for Hg,
but in my study >50% of the blood samples were BDL, which prevented inclusion in
analyses.
163
In Chapter 4, I conducted a similar contaminant uptake experiment where I
trapped American coots (Fulica americana) and ring-necked ducks from an
uncontaminated lake, clipped flight feathers on one wing, and released them on a
reservoir contaminated with radiocesium (Pond B) to model temporal patterns in uptake
of radiocesium between 32 to 173 days of exposure. Data from this experiment
corroborate a previous finding by Brisbin et al. (1989) that coots reached equilibrium
radiocesium burdens in ≤30 days of exposure. However, contrary to previous studies that
sampled free-ranging birds with unknown residence times, ring-necked ducks continued
to accumulate radiocesium up to approximately 90 days and had significantly higher
whole-body and muscle concentrations than coots. Furthermore, I observed a gradual
decline in radiocesium whole-body burdens after approximately 90 days of exposure,
suggesting potential seasonal differences in uptake due to changes in physiology or diet.
For both species, >90% sampled individuals had levels of radiocesium in muscle tissue
that exceeded the European Economic Community (EEC) limit for human consumption
of fresh meat, suggesting birds that spend only a few weeks on Pond B may accumulate
sufficient body burdens to exceed this threshold. Through the incorporation if historical
data for coots on Pond B, I was able to estimate an ecological half-life for radiocesium in
coots on Pond B of approximately 16.8 years (range of 12.9-24.2 years), suggesting it
will not be until approximately the year 2030 before the whole-body radiocesium in
coots, on average, would be below possible restrictive limits for human consumption.
An overarching theme of these three research experiments was that some game
species, and especially waterfowl, utilizing contaminated habitats on the SRS have the
capability to accumulate levels of trace elements and radiocesium that may be deleterious
164
to their own health and may pose consumption risks to human hunters. However, my data
clearly show how variability in residence time can affect the levels of accumulated
contaminants and thus it is imperative to incorporate this parameter into ecological risk
assessments whenever possible. Furthermore, animals that have access to contaminated
locations and have large home ranges (e.g. wild pigs) or are migratory (e.g.
waterfowl/waterbirds) should have special consideration when it comes to choosing focal
species for monitoring efforts, since these may act as vectors for transmission of
pollutants to humans far from containment sources.
The data presented herein build upon our understanding of contaminant
accumulation in wildlife and provide novel data on rates of contaminant uptake for
waterfowl inhabiting contaminated ecosystems. Given the high levels of several
contaminants observed in waterfowl/waterbirds in this study, including individuals that
likely accumulated these contaminants in non-SRS habitats, my data suggest a more
widespread effort is needed to better understand levels of contaminants found in game
bird muscle tissue and the potential risks (if any) to human hunters as well as the birds
themselves. Ultimately this information can help ecological lobbyists to petition for better
regulation of contaminated areas and managers to present consumption limits to the
public for common game species, like is regularly done for fish. My research results will
hopefully support wildlife conservation and management for these species that we are
stewards of and have an obligation to protect and propagate for future generations.