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RISK CHARACTERIZATION OF PERFLUOROALKYL ACIDS
EXPOSURE OF AQUATIC ORGANISMS IN LAKE SHIHWA,
KOREA
2
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(I) ._...
9E c :2
[I m g
CD .91: ' Hoon Yoo
_ c _
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has been accepted towards fulfillment
of the requirements for the

Ph.D. degree in Zoology and Program in
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Center for Integrative
Toxicolggy

 

 

 
  

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6/07 p:/CIRC/DateDue.indd-p.1

 

RISK CHARACTERIZATION OF PERFLUOROALKYL ACIDS EXPOSURE OF
AQUATIC ORGANISMS IN LAKE SHIW HA, KOREA

By

Hoon Yoo

A DISSERTATION

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY

Department of Zoology and
Program in Environmental Toxicology and
Center for Integrative Toxicology

2007

ABSTRACT

RISK CHARACTERIZATION OF PERFLUOROALKYL ACIDS EXPOSURE OF
AQUATIC ORGANISMS IN LAKE SHIWHA, KOREA

By

Hoon Yoo

The Shihwa-Banweol Industrial Complexes (SBIC), located on the west coast of Korea,
is one of the main national industrial complexes with a wide range of industries currently
operating. Recently, significant quantities of perfluoroalkyl acids (PFAs) were observed
in the waters of Lake Shihwa receiving wastewaters from SBIC. Thus, it was deemed
timely to determine concentrations of PFAs in aquatic animals of Lake Shihwa and assess
the potential risks that these compounds might pose to aquatic wildlife in Lake Shihwa

Aquatic samples (fish, blue crab, mussel, and oyster) and bird eggs were collected
in May to June 2006 from the Lake Shihwa area. All biotic samples contained
measurable concentrations of significant level of PFAs in their tissues. PFOS was the
predominant PFAs in fish species (mullet, rockfish, and shad), followed by the longer
chain perfluorocarboxylic acids (PFCAs), PFUnA > PFDA > PFDoA > PFNA > PFOA.
In the egg yolks of birds (little egret, little ringed plover, and parrot bill), measured
concentrations of PFOS were similar to birds eggs from other urban areas, but greater
than those from remote regions.

The spatial distribution of PFOS concentrations in marine organisms
demonstrates that biota samples from sites close to the outlets of inland creeks were more
contaminated than those sampled at sites away from the release sources. This

observation is consistent to the distribution of water-borne PFOS in Lake Shihwa. It

could be said that wastewaters from SBICs are at least one identified source of PFAs into
Lake Shihwa, consequentially contributing the elevated PFAs concentrations in the
marine wildlife.

In pharmacokinetic study, PFOS has a half-life of almost four months in male
chickens; in contrast, greater than half of introduced PFOA was eliminated within a week.
Thus, combined with a greater PFOS and a lesser PFOA in birds from the Lake Shihwa
area, at least current PFOA concentrations are unlikely to cause acute effects to birds.

For the hazard assessment of fish, PFOS body residues were compared to a
benchmark tissue concentration that would not be expected to cause acute effects in fish.
The calculated hazard quotients (HQs) were less than 1.0 for all species. Even a HQ
estimated from the greatest PFOS in fish was only to be 6X104. Thus, at least current
concentrations of PFOS in fish living in Lake Shihwa are not likely to cause acute
lethality.

Multiple lines of evidence were used to assess the PFAs associated-risks on birds,
which are one of the top-predators in the food web of the Lake Shihwa region. From a
bottom-up approach using fish as a sole diet for birds, a range of HQ (8.0><10'3 - 9.0X10’3)
generated was hundred-folds less than the least observable adverse effect level (LOAEL)
for PFOS. Similarly, the calculated HQs based on residue concentrations in egg yolk
were 8.0><10'3 for PFOS only and 9.0><10’3 for a mixture of PFAs, respectively, when the
LOAEL used as a benchmark dose. Although there are many uncertainties in deriving
these risk values, similar risk estimates from two opposite but complimentary approaches
indicate that current concentrations of PFOS and a mixture of PFAs would not be

expected to pose adverse effects to the avian population around the Lake Shihwa area.

To my parents, Hanoh Y00 and Even Shin, and my brother, Hyun Y00, and my wife,
Boram Yoo, for their love and support.

iv

ACKNOWLEDGEMENTS

The research contained in this dissertation would not have been possible without the
contributions of numerous top-quality scientists and people. Support and counsel were
provided by my graduate committee members, Drs. Thomas Burton and Patricia Ganey,
and Paul Jones. Dr. John P. Giesy, my graduate advisor, always provided encouragement
and support when I was in the darkness. Most importantly, he has showed me how to be
a good scientist. Drs. John L. Newsted and Daniel L. Villeneuve, and Markus Hecker
provided helpful comments and suggestions throughout this study. Collaborations with
scientists abroad were also integral to the success of this study. Drs. Yarnashita and
Guruge in Japan were invaluable to the completion of analytical aspects and animal
exposure study. I am also thankful to the graduate students at then I met while in Japan
for this study and helped me a lot; Drs. Taniyasu and Miyake from Japan, and MK. So
from Hong Kong, and Rostkowski from Poland. Scientist from Korea, Drs. Kyu Tae Lee
and Jong Hyeon Lee, and Tae Seob Choi (NeoEnbiz Co., Korea) and Lake Shihwa
wildlife expert Jong In Choi (City of Ansan, Korea), are thanked for their involvement in
aquatic sampling and egg collection around the Lake Shihwa area, Korea. This research
was supported, in part by a grant from the John P. and Susan E. Giesy Foundation to

Michigan State University.

TABLE OF CONTENTS

LIST OF TABLES ............................................................................. viii
LIST OF FIGURES .............................................................................. ix
LIST OF ABBREVIATIONS ................................................................ xi
CHAPTER 1
INTRODUCTION: PERFLUOROALKYL ACIDS ........................................... l
PFAs in the Environment ................................................................. l
PFAs as Emerging Contaminants ............................................... 1
Physical/Chemical Properties ................................................... 3
Sources to the Environment ................................................... 8
Environmental Transport ...................................................... lO
Toxicity of PF As .......................................................................... 13
Pharrnacokinetics and Toxicology .......................................... 13
Regulation of PFAs ............................................................ 16
PFAs in the Aquatic Environment ................................................ l9
Exposure Levels: Water, Sediment, and Aquatic Organisms .......... l9
Aquatic Toxicology: Algae, Invertebrates, Amphibians, and Fish 23
Study Design for Environmental Risk Characterization in Lake Shihwa ...... 27
The Lake Shihwa Area ...................................................... 27
PFAs Concentrations in Lake Shihwa Waters .............................. 29
Hazard Assessment Strategy ................................................ 32
CHAPTER 2

DEPURATION KINETICS AND TISSUE DISPOSITION OF PFOA AND PFOS IN
WHITE LEGHORN CHICKEN (G. gallus) ADMINISTERED BY SUBCUTANEOUS

IMPLANTATION ........................................................................... 36
Abstract ................................................................................. 36
Introduction ........................................................................... 37
Materials and Methods ................................................................ 40

Test Substances and Reagents ........................................... 40
Animal and Exposure ............................................................ 40
Sample Extraction ............................................................ 42
Matrix Recoveries .......................................................... 43
Instrumental Analysis and Data Analysis ................................. 43
Clinical Chemistry and Pathology ....................................... 44
Data and Statistical Analysis ............................................. 44
Results ................................................................................. 45
Body Index, Serum Biochemistry, and Histopathology ................ 45
Uptake Profiles and Elimination Behaviors in Blood ................. 49
Organ Concentrations ........................................................... 53
Discussion ........................................................................... 55

vi

PFOA vs. PFOS Elimination Kinetics ................................. 55

Organ Distribution ............................................................ 56
Sublethal Effects ............................................................ 62
Comparison with Other Studies ............................................. 63

CHAPTER 3
PERFLUOROALKYL ACIDS IN MARINE ORGANISMS FROM LAKE SHIHWA,
KOREA .......................................................................................... 65
Abstract .............................................................................. 65
Introduction ............................................................................... 66
Materials and Methods ............................................................... 68
Sample Collection ............................................................ 68
Chemicals and Standards ................................................... 70
Sample Preparation ......................................................... 70
PFAs Analysis by LC/MS/MS and QA/QC .............................. 71
Data and Statistical Analysis ................................................ 72
Results ................................................................................. 73
PFAs Concentrations in Fish .............................................. 73
PFAs Concentration in Marine Invertebrates .............................. 77
Profiles of Relative Concentrations of PFAa in Marine Organisms 80
Discussion ........................................................................... 82
Global Comparison of PFOS Concentrations ........................... 82
Source Estimation of PFAs in Lake Shihwa ........................... 85
BCFs for PFAs in Fish ...................................................... 87
Hazard Assessment Based on Concentrations of PF OS in Tissues ....... 89

CHAPTER 4
PERFLUROALKYL ACIDS IN THE EGG YOLK OF BIRDS FROM LAKE SHIHWA,
KOREA ........................................................................................ 91
Abstract .................................................................................. 91
Introduction ........................................................................... 92
Materials and Methods ................................................................... 95
Egg Collection and Sample Extraction ....................................... 95
Instrumental Analysis and QA/QC ............................................ 98
Gap Junction Intercellular Communication (GJIC) Cell Bioassay ........ 99
Statistical Analysis ............................................................. 100
Results and Discussion ............................................................ 101
Concentrations in Birds ........................................................ 101
Contamination Sources to Birds and Relationship among PFCAs ...... 105
PF OS-Equivalent Concentration ............................................. 108
Risk Characterization of PF OS and Other PFAs ........................ 112
CONLCUSION ................................................................................... 1 1.7
REFERENCES .............................................................................. l 18

vii

LIST OF TABLES

Table 1. Physical/chemical properties of PFOS and PFOA .................................. 6
Table 2. Persistence parameters available for PF OS and PFOA ............................. 7
Table 3. Concentrations of PF OS and PFOA in waters (ng/L) .............................. 20
Table 4. Mean concentrations of perfluoroalkyl acids in aquatic biota (ng/ g, wet wt. for
liver and ng/mL for blood) ........................................................................ 22
Table 5. Toxicity to aquatic organisms (mg/L) ............................................... 24

Table 6. Concentrations of PFAs in water samples from streams, Lake Shihwa, and
Gyeonggi Bay, Korea .............................................................................. 31

Table 7. Mean body-weight gain (s.d.) and organ to body weight ratio at the end of an
exposure (A) and an elimination period (B) ................................................... 47

Table 8. Mean serum lipids and biochemistry measurements for experimental chickens
determined at the end of an exposure (A) and an elimination period (B) (n=3) (*2 p<0.05)
......................................................................................................... 48

Table 9. Elimination half-life (days) of PFOA and PFOS in blood or serum from other
studies ............................................................................................... 64

Table 10. Mean PFAs concentrations (ng/g wet wt.) in the tissues of fish collected from

Lake Shihwa, Korea ............................................................................... 75
Table 11. PFAs concentrations (ng/g wet wt.) in the soft tissues of marine invertebrates
collected from Lake Shihwa, Korea ............................................................. 79
Table 12. Birds egg sample descriptions ...................................................... 97
Table 13. Concentrations of perfluoroalkyl acids in the egg yolks of birds collected
around Lake Shihwa (ng/ g wet wt.) ............................................................ 102
Table 14. ECSO of PFAs and toxicity equivalent factor (TEF) ........................... 112

viii

LIST OF FIGURES

Figure 1. Structure of PFAs of current interests ............................................. 5
Figure 2. The water sampling locations in Lake Shihwa areas ............................. 28

Figure 3. Dissertation research scheme and assigned works in each chapter
......................................................................................................... 33

Figure 4. Uptake profiles of PFOA and PFOS (ng/mL) introduced into male chickens
using an implantation for a 4-wk exposure period (n=2) .................................... 51

Figure 5. Blood concentrations of PF 0A or PF OS (ng/mL) in male chickens over a 4-wk
depuration period (n=2) ........................................................................... 52

Figure 6. Concentrations (ng/g wet wt.) of PFOA (A~C) or PFOS (D~F) in organs
retrieved from chickens at the completion of an exposure and a depuration period (n=3)
......................................................................................................... 54

Figure 7. Normalization of individual organ concentration to blood concentration after
an exposure (A) and a depuration (B) ........................................................... 59

Figure 8. Relative mass (%) of PFOA or PFOS in body reservoirs (blood, brain, kidney,
and liver) after an exposure and a depuration period ....................................... 60

Figure 9. Map showing sampling locations in Lake Shihwa study for marine fish and
invertebrates ......................................................................................... 69

Figure 10. Species comparison of PFA concentrations in mullet, shad, and rockfish

collected from Lake Shihwa ..................................................................... 76
Figure 11. Composition profiles of PFAs in liver and blood of mullet, shad, and rockfish
collected from Lake Shihwa ...................................................................... 81
Figure 12. Comparison of PFOS concentrations in liver (ng/g wet wt.) and blood
(ng/mL) of fishes from Japan, Lake Shihwa, and other global locations .................. 84
Figure 13. Plot of the odd-chain PFCA and the even-chain PFCA in liver (A) and blood
(B) of fish collected from Lake Shihwa ......................................................... 88
Figure 14. Map showing sampling locations for birds eggs in this study ............... 94

Figure 15. Species comparison of PFOS and total PFCAs concentrations (ng/g wet wt.)
in little egret, little ringed plover, and parrot bill from the Lake Shihwa area
....................................................................................................... 103

ix

Figure 16. Composition profiles of PFAs in the egg yolks of little egret, little ringed
plover, and parrot bill collected from Lake Shihwa and its vicinity ....................... 107 _

Figure 17. Plot of the odd-chain PFCA and the even-chain PFCA in the egg yolks of
birds from Lake Shihwa and its vicinity ...................................................... 109

Figure 18. Concentration-response relationship between inhibition of cellular
communication and a gradient of four PFAs (PFHS, PFOS, PFOA, and PFDA) on rat
liver epithelial cells .................................................................... 111

Figure 19. Cumulative percent rank against concentrations of PF OS and PFOS-EQ in the
egg yolks of birds from Lake Shihwa and its vicinity .................................... 114

LIST OF ABBREVIATIONS

PF A: Perfluoroalkyl acid

SBIC: Shihwa-Banweol Industrial Complex
PFOS: Perfluorooctanesulfonic acid

PFHS: Perfluorohexanesulfonic acid

PFAS: Perfluoroalkyl sulfonates

FOSA: Perfluoroalkyl sulfonamide

PFCA: Perfluorocarboxylic acid

PFDoA: Perfluorododecanoic acid

PFUnA: Perfluoroundecanoic acid

PFDA: Perfluorodecanoic acid

PFNA: Perfluorononanoic acid

PFOA: Perfluorooctanoic acid

PFO: Perfluorooctanoate

PFHA: Perfluorohexanoic acid

F TOH: F luorotelomer alcohol

BCF: Bioconcentration factor

HQ: Hazard quotient

TRV: Toxicity reference value

GJIC: Gap junction intercellular communication

LOAEL: Lowest observable adverse effect level

xi

Chapter 1

Introduction: Perfluoroalkyl acids

PFAs in the Environment

PFAs as emerging contaminants

Perfluoroalkyl acids (PFAs) are fully fluorinated synthetic compounds that have been
used for a wide range of industrial applications and in the manufacture of commercial
products since their introduction in the late 1940’s. PFAs are used in the production of
refrigerants, surfactants, polymers, pharmaceuticals, wetting agents, pesticides, and fire-
fighting formulations (Kissa, 2001). The popularity of PFAs in industrial areas comes
from their favorable physico-chemical properties, such as strong
chemical/biological/therma1 resistance, special surface-activity, and both lipid- and
water-repellency. However, the scientific community became widely aware that there
could be potential environmental problems with PFAs when one of the major
manufactures, 3M, announced the voluntary phase-out of perfluorooctanesulfonyl
fluoride-based products from the market in 2000 (Renner, 2001). At that time, there was
limited information on the toxicity of the PFAs and due to a lack of standards and
methods, even less information on the concentrations of these compounds in the
environment. Since then, more toxicological tests have been conducted and revealed that
PFAs, which were once considered to be biologically inert, are bioactive at the cellular
level and are toxic to laboratory animals. Increasing numbers of peer-reviewed papers
have pinpointed the potential environmental hazards posed by PFAs. Drinking water

contamination near DuPont’s Washington Works facility in West Virginia made the

public aware of these emerging contaminants (Hogue, 2005). Among the family of PF As
and related compounds, two eight-carbon PFAs, perfluorooctanesulfonate (PFOS) and
perfluorooctanoic acid (PFOA), have received much of attention from environmental
scientists, toxicologists, and legislators than other PFAs. These two compounds are
considered to be the ultimate degradation products of PFAs and related compounds, and
exposure levels are great in the environment (Giesy and Kannan, 2002; OECD, 2002; US
EPA, 2003). PFOS was the primary active ingredient in 3M Scotchgard® stain repellent,
and PFOA is used in DuPont’s Teflon® products.

PFAs are ubiquitous global contaminants that are considered to be PBT
(persistent, bioaccumulative, and toxic) contaminants, a category which also includes
compounds such as DDTs, PCBs, and dioxins. Their presence has been confirmed in the
serum or blood of the general public from various countries, different ethnicities, and all
ages (Kannan et al., 2004). They have also been found in wildlife in remote regions of
the globe such as Arctic marine mammals (Giesy and Kannan, 2001; Bossi et al., 2005).
Food web studies and laboratory experiments have shown that PFAs are bioaccumulative
to some degree and able to biomagnify in fish-eating top predators, such as mink, eagles,
and polar bears (Martine et al., 2003a & 2003b; Giesy et al., 2006). Generally,
monitoring has shown that PFA levels in abiotic matrices, wildlife, and humans are
higher in urban and industrialized areas than in rural and less industrialized areas. This
observation implies that, although they are globally ubiquitous, releases of PFAs into the
environment are closely related to human activities. As a consequence, legal measures

and management programs have been initiated and developed in order to reduce the

environmental release of PFAs (See the section for Regulation of PFAs). Replacements
for PFOS- or PFOA-based chemicals are already on the market.

At present, it is quite challenging to analyze the risks of PFA exposure to humans
and wildlife. First of all, now there are vast numbers of known PFAs and unidentified
fluorinated chemicals used in the world (Kissa, 2001). Analytical chemists are still trying
to develop new technology such as efficient and reliable analytical methods and
instruments (Martin et al, 2004a). Although sufficient toxicological data are available for
mammals, toxicologists have not yet pinpointed the mechanisms underlying the toxic
responses. Data generated prior to development of proper analytical techniques, are
somewhat unreliable (Susan et al., 2006). Most toxicological studies have investigated
short-term exposure scenarios and may not be reliable for predicting long-term effects.
Distinct pharmacokinetic characteristics between humans and other mammals also
prevent risk assessors from extrapolating animal-based outcomes. The contamination
sources and environmental fates of PFAs are not fully identified with few proposed

hypotheses, which make it difficult to properly manage exposures to living organisms.

Physical/chemical properties

In PFAs or poly-fluorinated chemicals, a number of fluorine atoms are substituted for
hydrogen atoms along their carbon backbone. This replacement with strongly
electronegative fluorine brings unique chemical properties to PFAs, such as lowering pKa.
The overlapping of the 28 and 2p orbitals of fluorine with the corresponding orbitals of
carbon results in effective shielding of carbon chains, and contributes to the rigidity of

perfluorocarbon chains (Key, 1997; Kissa, 2001). Generally, the structure of PFAs

consists of a non-polar perfluoroalkyl tail, and an ionic or neutral functional group
(Figure l).

The tail of PFAs is characterized by both oleophobic (oil-repelling) and
hydrophobic (water-repelling) properties, which are in contrast to the hydrophobic
hydrocarbon chain. Commonly, PFAs with sulfonic and carboxylic group are called
perfluoroalkyl sulfonates (PFASs) and perfluoroalkyl carboxylates (PFCAs), respectively.
PFAs with charged moieties are better surfactants than hydrocarbon-based surfactants
(Giesy and Kannan, 2002). PFAs can lessen the surface tension of aqueous solutions and
lower critical micelle concentrations more than hydrocarbon-based surfactants. In
addition, these highly polarized, high-energy C-F bonds are thermodynamically stable;
thus they are largely resistant to metabolism, microbial degradation, hydrolysis, and
photolysis (OECD, 2002). Strong chemical stability to acids, alkalis, and oxidizing
agents also makes them useful in operating conditions where hydrocarbon-based
surfactants cannot be used, such as in metal plating or in fire extinguishers (UN EP, 2006).
Plausible precursor molecules of PFASs and PFCAs are also gaining some interests,

including perfluoroalkyl sulfonamides (FOSAs) and fluorotelomer alcohols (FTOHs).

PFASs
(Perfluoroalkyl sulfonates)

ii
F(CF2)n 7s — OH
0

PFDS (Perfluorodecanesulfonate)
PFOS (Perfluorooctanesulfonate)
PFHxS (Perfluorohexanesulfonate)

' FOSAs
(Perfluoroalkyl sulfonamides)

o R

ll /
F(CF2)n —s- N

“ \

o R

PFOSA (Perfluorooctanesulfonamide)
N-EtFOSA (N-ethyl FOSA)
N-MeFOSA (N-methyl FOSA)
N-EtFOSEA

N-MeFOSEA

N-EtFOSE alcohol

N-MeFOSE alcohol

PFCAs
(Perfluoroalkyl carboxylates)

ii
F(CF2)n -c- OH

PFDoDA (Perfluorododecanoic acid)
PFUnDA (Perfluoroundecanoic acid)
PFDA (Perfluorodecanoic acid)
PFNA (Perfluorononanoic acid)
PFOA (Perfluorooctanoic acid)

FTOHs
(Fluorotelomer alcohols)

F(CF2)n — (CH2)2 — OH

6:2 FTOI-l
8:2 FTOH
10:2 FTOH

Figure 1. Structure of perfluorinated acids (PFAs) of current interests.

The physico-chemical properties and persistence parameters of PFOA and PFOS
are summarized in Table 1&2. However, these estimates should be used with caution,
since they were not measured under well-controlled experimental conditions, and many
essential data are still lacking. PFOS is extremely persistent. PFOS does not hydrolyze,
photolyze or biodegrade in any environmental condition tested (OECD, 2002). It is not
expected to volatilize based on an air/water partitioning coefficient of <2><10’6 Pa m3/mol
and it was classified as a type 2, nonvolatile chemical by OECD. PFOS is moderately

water soluble (680 mg/L in pure water).

Table 1. Physical/chemical properties of PFOS and PFOA "

 

 

Parameter PFOS . PFOA
(Potassrum salt)

Melting point 2400 °C 45-50 °C

Boiling point Not calculable 189-192 °C at 736mm Hg

Specific gravity b ~ 0.6 (7-8) —

Vapor pressure 3.31 X lOAPa at 20 °C 1.33><10'5 Pa at 25 °C
(PFOA ammonium salt)
1.33><104 Pa at 25 °c (PFOA)

Water solubility

Pure water 680mg/L at 20°C 3.4g/L
Fresh water 370mg/L
Sea water 12.4mg/L

Octanol solubility 56mg/L

Log K...“ -l.08

Henry’s law constant d 4.34><10'7

(atm.m3/mol)

 

“ Data were collected from OECD (2000) and Giesy et a1. (2006)

a pH values in parentheses

b Log Kow calculated with solubility of PFOS in water and n-octanol

° Henry’s law constant calculated at 20 with the solubility for pure water

Table 2. Persistence parameters available for PFOS and PFOA

 

 

PF A Media Study Degradation half-life Reference
PFOS Water Hydrolysis . 2 41yr at 25°C a 3M Company 2001b
Activated- Biodegradation No loss after 20 wk Gledhill&Markely,2000
sludge
Photolysis Z 3.7yr at 25°C ° 3M Company 2001a
PFOA Water Photolysis >349 days Hatfield, 2001
Water Photolysis No loss after 35 days Oakes et al., 2004
Hydrolysis
Biodegradation ~235 days
Water Hydrolysis ~90.1 days US EPA, 2002
OH Reaction Hurley et al., 2004

 

a This estimate was influenced by the analytical limit of quantification and that no loss of
PFOS was detected in the study.

b No evidence of direct or indirect photolysis of PF OS in the study. The indirect
photolytic half-life was estimated using an iron oxide photoinitiator matrix model

However, the solubility of PFOS decreases with increasing dissolved solids as a result of
a salting-out effect. Available data indicate PFOA also does not significantly photolyze,
hydrolyze, or undergo abiotic or biotic degradation under environmental conditions. The
solubility of PFOA is reported to be as much as 3.4 g/L. PFOA as the free acid is capable
of escaping from water to air; however, this is unlikely to be a significant process due to
the strong acidity of PF 0A with a pKa of ~2.5, thereby the virtually non-volatile
conjugate base (PFO, C3F1502') of PF OA acid (C3HF1502') is dominant at environmental
conditions (Prevedouros et al., 2006). Accurate prediction of the environmental fate and
transport of PFAs is very difficult due to the lack of accurate physico-chemical

parameters for most PFAs. For example, the octanol-water partitioning coefficient (Kow)

is commonly used to describe the bioavailability of organochlorine contaminants (e.g.,
PCBs, dioxins, and organochlorine pesticides). In biotic samples, however, PFAs are
often accumulated in blood and organs related to enterohepatic circulation rather than in
lipids. This behavior is attributed to the lipophobic and hydrophobic properties of PFAs,
which form three immiscible layers when they are experimentally added to octanol (a
lipid surrogate) and water. Thus, the utility of estimated K0W values for PFAs is

questionable.

Sources to the environment

There are both direct and indirect sources of PF As emission to the environment. Direct
sources include the manufacture and application of PFAs, while indirect sources result
from the liberation of impurities in finished goods containing PFAs or their precursors
which can be degraded to the final metabolites, PFOS and PFOA (UNEP, 2006;
Prevedouros et al., 2006).

Direct sources - During manufacturing processes, PFAs may be discharged in
sewage effluent and the volatile precursors of PFSAs and PFCAs may be released to the
atmosphere. Currently, these manufacturing processes are thought to be a major source
of PFAs to the local environment. For example, the estimated emissions of PFCAs via
direct sources accounted for almost ~80% (3.2><1O°-6.9><103 tones) of total global
emissions, and fluoropolymer manufacturing facilities were the single largest known
source (2.4><103-5.4><103 tones) of PFCA emissions (Prevedouros et al., 2006). As a
consequence, effluents from fluorochemical manufacturing facilities have elevated

concentrations of perfluoroalkylated substances. Concentrations of PFOS and PFOA

measured downstream of 3M’s fluorochemical manufacturing facility in Decatur, AL,
were as great as 0.11 rig/L and 0.39 rig/L, respectively (Hansen et al., 2002), while these
PFAs were not detected upstream of the facility. A variety of different uses of PF As are
reported in the surface treatment, paper protection, and performance chemicals. The
PFOS concentration in effluents collected from representative industry in Australia
ranged from 0 to 2.5 rig/L (2.5 rig/L for leather, 0.12 pg/L for metal, O.14-1.2 rig/L for
paper, 1.2 pg/L for photographic, not found in textiles or electronics) (Hohenblum et al.,
2003). PFAs are present at elevated concentrations, not only in the industrial
wastewaters, but also in public wastewater. In a multi-city study by the 3M Company,
PFA concentrations in publicly-owned treatment works effluents ranged of 0.05-4.98X102
ug/L for PFOS and 0.04-2.28 ug/L for PFOA (3M, 2001). The occurrence of PFAs in
corresponding sludge samples was also observed in that study. Application of
commercial products containing PFAs, such as fire-fighting foams, is another source to
the environment. Aqueous film-forming foams (AFFF), which are used to extinguish
liquid-fuel fires, contain PFAs as a key ingredient. As a result of historic fire-fighting
training exercises at Air-Force Bases, elevated levels of PFOS (8 to 1.10><102 pg/L) and
PFOA (not detected to 1.05>< 102 ug/L) have often been detected in groundwater in the US
close to fire-training areas (Moody etal., 2003).

Indirect sources — Liberation of PFSA or PF CA residual impurities in products as
well as precursor degradation are assumed to increase the levels of PFOS and PFOA in
the environment, although these processes are of lesser importance than direct sources.
For example, fluorotelomer-based products may contain trace levels of PFCAs (<l-100

ppm) as unintended reaction byproducts (Prevedouros et al., 2006). Significant amount

of precursor molecules unbound to fluorinated materials have been measured in
commercially available carpet and rug protectors (Joyce et al., 2006). For example, air
samples from Griffin, GA, USA had the greatest airborne FOSAs and FTOHs
concentrations, perhaps, because Griffin is located in the midst of western Georgia and
eastern Alabama, a major carpet manufacturing and treatment zone. Indoor air
concentrations of FOSAs were between 10 and 20 fold greater than outdoor
concentrations in Canadian houses (Shoeib et al., 2005). Laboratory experiments have
suggested that these precursor molecules can be broken down into PF OS and PFOA via
biotic and abiotic pathways (See the section for Environmental transport). Both F OSAs
and FTOHs, which carry a PFOS and a PFOA moiety, respectively, are widely

distributed throughout the North American troposphere (Stock et a1 ., 2004).

Environmental transport

The ubiquitous occurrence of the least volatile PFAs at often elevated levels in Arctic
mammals has been an interesting issue, in that PFOS and PFO (perfluorooctanoate, the
dominant form of PFOA at environmentally relevant pH) have a relatively low tendency
to partition to the atmosphere and undergo long-range transport (Giesy and Kannan,
2002). Two transport mechanisms have been proposed to explain this unexpected
observation. Atmospheric chemists hypothesized that more volatile precursors can be
transported to the Arctic region by air currents and subsequently degraded to form PF OS
and PFCAs via atmospheric reactions or metabolism (Ellis et al., 2004; Wallington et al.,
2006). For example, the vapor pressures of PFOS precursors (ca 0.5 Pa for N-EtFOSEA

and N-MeFOSEA) are almost 1000 folds greater than PFOS (Giesy and Kannan, 2002).

10

The relatively great solubility of PFAs in water, in particular of PFCAs, supports an
aquatic transport route to the Arctic. The development of more accurate analytical
techniques to analyze trace levels of PFAs allowed the determination of PFAs in open
oceans, indicating the oceanic movement of certain PFAs to remote regions via major
currents. Seawater from the Atlantic and Pacific Ocean were contaminated with trace but
measurable amounts of PFAs (pg/L) such as PFOS, PFHS, PFOA, and PFOSA
(Yamashita et al., 2004). The quantity of PFO transported through the ocean currents
flowing to the Arctic was estimated to be 2-12 tones per year; this route contributes a
greater amount of PF As to the Arctic than the estimated atmospheric transport route (0.1-
] ton/year) in the northern hemisphere from FTOH degradation and subsequent
atmospheric deposition (Prevedouros et al., 2006).

Several studies have supported the widespread global presence of the least
volatile PFAs using their semi-volatile precursors. The non-fluorinated parts of precursor
compounds, such as FOSAs and FTOHs can be degraded into PFOS and PFCAs via
atmospheric reactions, metabolism, and microbial activity (Ellis et al., 2004; Tomy et al.,
2004; Dinglasan et al., 2004). Of those mechanisms, atmospheric dispersion and
degradation of F TOHs is thought to be a pathway for the global dissemination of PFCAs.
In a smog chamber experiment, chlorine atom-initiated oxidation of FTOHs leads to the
formation of small, but significant yields of PFCAs (Ellis et al., 2004). In their
experiment, the addition of 8:2 FTOH generated almost 4% of total PFCAs in the
absence of NOx in the air. Hydrogen atom abstraction by OH radicals from the —CH2-
group is suspected to be an initial reaction of FTOHs oxidation. Laboratory incubation

tests showed that N—ethyl perfluorooctane sulfonamide (N—EtFOSA or a Sulfluramid®,

ll

which is used as an insecticide) was biotransformed by rainbow trout (Onchorhynchus
mykiss) into PFOS over time (Tomy et al., 2004). In metabolic products of FTOHs in rat
hepatocytes, PFOA and minor amounts of PFN A were confirmed. Similarly, in a mixed
microbial system, FTOHs biodegradation yielded PFOA and telomer acids (Dinglasan et
al., 2004). The authors hypothesized that the main mechanism was the oxidation of the
8:2 FTOH to the telomer acid via the transient telomer aldehyde followed by the ultimate
formation of the stable PFOA via a B-oxidation of the telomer acid.

The well-established equilibrium partitioning (Eq-P) theory predicts the phase-
partitioning behavior of nonpolar organic contaminants among water, biota, and
sediments in the aquatic environment (Di Toro et al., 1991). Hydrophobic interactions
and Kow are a key theoretical basis of this concept, which has been employed to set
environmental criteria for non-polar chemicals (e.g., PAHs and chlorinated organic
chemicals). However, the perfluorinated tail of PFAs is not only hydrophobic, but also
oleophobic (Kissa, 2001). Thus, Eq-P theory is inadequate to describe the environmental
fate of PFAs (Giesy et al., 2006). Most commercial perfluoro-surfactants are synthesized
as salts and have intrinsically polar natures. As a result, they often exist as dissociated
salts in aqueous environments rather than as the acid form at environmentally realistic pH.
Therefore, ionic and electrostatic interactions between compartments may be important to

describe environmental behaviors of PFAs discharged into the aquatic environment.

12

Toxicity of PFAs

Pharmacokinetics and toxicology

PFAs have distinct elimination kinetics and tissue disposition patterns compared to other
persistent organic pollutants (POPS). Notably, they are bound to serum proteins such as
albumin and lipoproteins, and are preferentially accumulated in the liver and gall bladder,
not in fatty adipose cells (Luebker et al., 2002; Jones et al., 2003; Martin et al., 2003a).
Therefore, the target site of action would be the liver of an exposed receptor organism.
The abilities to eliminate PFAs from the body are very different among species and sex.
Epidemiological studies and medical monitoring of retired fluorochemical workers reveal
that PFOS and PFOA are rarely eliminated from the body without notable sex differences.
For example, the biological half-life of PFOS and PFOA in occupationally-exposed
people is estimated to be 1428 and 1600 (1, respectively (3M, 2000). However, salient
differences in elimination kinetics of PFOS and PFOA have been found between males
and females, and across experimental animals (Butenhoff et al., 2004a). In general, both
chemicals are quickly absorbed, but PFOS tends to be more persistent in the body than
PFOA. For example, the elimination half—life for PF OS in male rats was more than 90 (1,
while that of PFOA was 4.4 to 9 d (Lau et al., 2004). Sex differences are also obvious in
some animals. For example, half of the blood-home PFOA in female rat was removed
within 24 hr after an oral dose, compared to 6 to 8 d for male rat (Kemper, 2003). This
gender difference is also seen in exposed dogs. So far, scientists have found that organic
anion transporters (OATS), which are regulated by sex hormones, play an important role
in depurating PFAs, consequently resulting in different renal clearance rates between

females and males. For example, OAT2 activity in the kidney of adult male rats is only

13

13% of that of female rats (Kudo et al., 2002). The amount of OAT2 can be increased in
the kidney of males by treatment with estradiol or castration.

With respect to their toxicology, mammalian studies have reported developmental
and reproductive toxicities (Kennedy et al., 2004; Lau et al., 2004; Betts, 2007). The
ammonium salt of perfluorooctanoate (APFO) and other PFAs (e.g., PFOS, and
perfluorodecanoic acid (PF DA)) are well known peroxisome proliferators and inducers of
hepatic CYP450 enzymes. The hepatotoxicity of PFAs is reversible on termination of
dosing, and manifests as increased liver weight due to hepatocellular hypertrophy and
proliferation of smooth endoplasmic reticulum and peroxisomes, increases in plasma
transaminases, liver degeneration and necrosis in rodents (Butenhoff et al., 2004b;
Luebker et al., 2005). Exposure to PFOS and N—ethyl-N-(2-hydroxyethly)-
perfluorooctanesulfonate (which is converted metabolically to PFOS) impaired
reproductive system function in rats, causing an increase in perinatal mortality (Lau et al.,
2003). Neonatal mortality was also observed following PFOS exposure in a two-
generation reproduction study in rats (Luebker et al., 2005). PFOS did not affect
reproductive performance (mating, estrous cycling, and fertility), except at higher doses
(3.2mg/kg/day) where impaired reproductive outcomes were found, including decreased
length of gestation and number of implantation sites, and an increased number of stillborn
pups. The NOAEL for offspring effects was 0.4mg/kg/day (oral dosing) based on
mortality and decreased pup weight gain through lactation observed in the F1 generation.

In contrast to PFOS, the reproductive system was not the primary target of PFOA
in a rat study. Reproductive performance was unaltered by dosing up to 30 mg/kg APFO,

including fertility, sperm number and quality, number of liveborn/stillbom pups,

l4

gestation lengths, and estrous cycling (Butenhoff et al., 2004b). No evidence of embryo-
fetal toxicity or gross developmental abnormalities was noted in the studies that evaluated
the oral and inhalation routes of APFO in rats, even at the doses that produced maternal
toxicity (Stapels et al., 1984). Changes in body weight, which is the most sensitive
indicator of APFO exposure, showed an interesting outcome across multiple phases of
sexual maturation and gender. Decreased pup weights, increased pup mortality, and
delayed sexual maturation in Fl-generation offspring were seen at a higher dose (30
mg/kg) in rats. Sprague-Dawley male rats of either the parent or F-l generation exposed
orally to APFO showed statistically lower body weight compared to vehicle control rats,
but this reduction was not observed in female rats. In addition, weight decrements of
sexually mature male rats (%-control) were more severe than sexually immature animals
at a given concentration.

PFAs are a known surface-active agents and are structurally analogous to
endogenous fatty acids (e.g., hormones), meaning that they are potentially toxic at (sub)-
cellular levels. Studies have shown that they are inhibitors of gap junction intercellular
communication (GJIC) and disruptors of membrane integrity (Yoo et al., 2005; Hu et al.,
2002). Cellular communication through gap junctions is critical to maintain cellular
homeostasis, growth, differentiation, and apoptosis. Interference with cellular
communication may lead to disease and disruption of cellular integrity. PF OS and PF 0A
are known to be potent PPAR-a agonists and to increase peroxisome proliferation in adult
rodents (Ikeda et al., 1985; Seacat et al., 2003). Microarray analysis suggested that PFAs
can modulate gene expression, and provided insight into the effects of PFAs on many

biochemical pathways (Hu et al., 2005). Exposure of Sprague-Dawley rats to PFOA and

15

PFOS resulted in the common up-regulation of genes involved in fatty acid and lipid
metabolism such as peroxisomal fatty acid beta-oxidation. PFOA treatment suppressed
genes related to lipid transport, inflammation, immunity, and especially cell adhesion,
while PFOS treatment down-regulated. genes involved in signal transduction and

neurosystem regulation.

Regulation of PFAs

In a recent risk assessment aimed for the public exposure to PFOA that is one of the
predominant PFAs in human and wildlife, the margin of exposure (MOE) was estimated
to range from 1600 (liver-weight increase) to 8900 (Leydig cell adenoma), and the
assessment concluded that the MOE values represented substantial protection of children,
adults, and the elderly at the present exposure levels (Butenhoff et al., 2004a). This study
derived a conservative MOE value using multiple biological responses from broad
toxicokinetic and toxicodynamic studies by comparison of serum PFOA concentrations in
a general or non-occupationally exposed population.

In addition to voluntary actions by PFA manufactures to stop or minimize the
release of these chemicals, especially with regard PFOS and PFOA, into the environment,
some legislative measures and management programs are now being developed in
countries including US, Canada, and UK, and in international organizations including
OECD, EU, and UNEP.

The US EPA finalized two Significant New Use Rules (SNURs) in 2002,
requiring companies to inform the EPA before manufacturing or importing 88 listed

PFOS-related substances with the exception of a few chemicals with essential uses in

16

aviation, photography, and microelectronics (US EPA, 2002). Four years after the initial
withdrawal of PFOS-based products, the EPA proposed an additional SNUR under the
Toxic Substances Control Act (TSCA) in March 2006, which included another 183
perfluoroalkyl sulfonates with carbon chain lengths of five or more carbons in regulation
(US EPA, 2006). EPA further proposed an amendment to the Polymer Exemption rule in
2006 which would remove polymers containing perfluoroalkyl moieties (CF3-) or longer
chains from exemption.

A preliminary draft Hazard Assessment of PFOA released by the US EPA in
August 2002 found that PF 0A and its salts are of similar concern due to their structural
analogy with PFOS. Currently, human health effects of long-term exposure to PFOA and
its salts are the subject of an EPA risk assessment. The EPA is taking action to help
minimize the potential impact of PFOA on the environment. In January 2006, EPA
initiated the PFOA Stewardship Program, in which the eight major companies in the
fluoropolymer and telomere industry committed voluntarily to reduce facility emissions
and product content of PFOA and related chemicals on a global basis by 95% no later
than 2010, and to work toward eliminating emissions and product content of these
chemicals by 2015. Participants in this program include, 3M/Dyneon, Arkema,Inc., AGC
Chemicals/Asahi Glass, Ciba Specialty Chemicals, Clariant Corporation, Daikin, E.I.
duPont de Nemours and Company, and Solvay Solexis.

The final Environmental Canada/Health Canada assessments of PFOS, its salts
and its precursors were released in July 2006. The ecological risk assessment has
concluded that PFOS and its salts are persistent and bioaccumulative. PF OS, its salts and

its precursors have immediate or long-term harmful effects on the environment

17

(Environmental Canada, 2006b).

The hazard assessment of PFOS prepared by the OECD in 2002 following the
34th Joint Meeting endorsed a draft assessment of PF OS and its salts by the United States
(US) and the United Kingdom (UK). The Hazard Assessment concludes that PFOS and
its salts are persistent, bioaccumulative and toxic to mammalian species. The US EPA
made a request for an international measure to be passed on PFOS, but the OECD did not
recommend action for a ban (UNEP, 2006). Instead, they recommended that
governments contact PFOS manufactures in their country to determine whether the
companies have plans to stop PFOS production. In 2003, OECD also sent a
questionnaire to OECD member and non-member countries requesting information on the
production and use of PFOS, PFOA and perfluoroalkyl sulfonates and products/mixtures
containing these chemicals.

An environmental risk assessment prepared by the UK—Environmental Agency
and discussed by the EU member states under the umbrella of the existing substances
regulation (ESR DIR 793/93) shows that PFOS is of concern (UNEP, 2006). In June
2004, the UK government announced unilateral action to phase out PFOS and related
compounds because of their persistence and bioaccumulative potential. The EU has
recently decided on restrictions on the marketing and use of PFOS (EU, 2006). The
measures cover PFOS acid, its salts and PFOS derivatives, including PFOS polymers.
The decision prohibits the sale and use of these compounds as a substance or constituent
of preparations in a concentration equal to or higher than 0.005% by mass. Furthermore,
semi-finished products and articles containing PFOS at more than 0.1% by mass are

prohibited

18

PFAs in the Aquatic Environment

Exposure levels: Water and aquatic organisms

Chemodynamic values of PFOA and PFOS indicate they are likely to exist in the
dissolved phase rather than in the gaseous phase (Table 1&2). For example, the low
Henry’s Law constants of PFOA and PFOS estimated indirectly from high water
solubility and negligible volatility predict that all PFOA and PFOS species are expected
to partition primarily to water. PFOA is the major contaminant measured in water
samples from freshwater to seawater (Table 3). In general, levels of PFAs in inland
waters were greater than those in coastal waters. Other PFAs such as PFHS and PFNA
are also present, but often at insignificant levels (data not shown). PF OS concentrations
in the Great Lakes ranged from 6 to 1.2><102 ng/L for Lake Ontario and from 1.1><10l to

3.98101 ng/L for Lake Erie.

19

Table 3. Concentrations of PFOS and PFOA in waters (ng/L)

 

Freshwater PFOS PF OA Ref.
Lake Ontario, USA/Canada 6-121 15-70 Boulanger et al., 2003
Lake Erie, USA/Canada 11-39 21-47 Boulanger et al., 2003
Tennessee River, USA 17-144 <25-598 Hansen et al., 2002
Pearl River, China [-94 <1-l3 So et al., 2007
Yangtze River, China <1-13 2-245 80 et al., 2007
Effluent Water
New York, USA 3-68 58-1,050 Sinclair, 2006b
Coastal Water
Korea <0.1-3 0.2-11 So et al., 2004
Tokyo Bay, Japan 0.3-58 1.8-192 Yamashita et al., 2004
Hong Kong, China <0.1-3 0.2-5 So et al., 2004
Nordic seawater, Norway <0.3-22 3.5-8.5
Open Ocean
Atlantic Ocean <0.l 0.1-0.4 Yamashita et al., 2004
Pacific Ocean <0.] <0. 1-0.1 Yamashita et al., 2004

 

A mean concentration of PFOS upstream of a fluorochemcial manufacturing facility in
Tennessee River, AL, USA was 3.2><10l ng/L; however, PFOS concentrations were
almost four-fold greater downstream of the plant, with a mean concentration of 1.1><102
ng/L. PFOA showed a contamination profile similar to PFOS in this river. In the
Tennessee River, the elevated levels measured did not change downstream of the facility,
suggesting the absence of physical removal mechanisms from water, such as
volatilization or adsorption to soil or sediment. In effluent collected from wastewater
treatment plants in New York, PFOA concentrations were as great as 1.05><103 ng/L, but

the greatest PF OS concentration measured was 6.8><10l ng/L. PFCAs with a chain length

of 9 to 11 carbons were also present in effluents. In coastal waters, Tokyo Bay was more

20

contaminated with PFAs than other East Asian coastal waters. Surprisingly, trace but
measurable levels of PFAs were detected in open oceans. Interestingly, sediments play a
less important role in PFA accumulation than water. While other persistent organic
contaminants often occur at the parts-per million level in sediment, sedimentary PFA
concentrations are at the low parts-per-billion level.

In a contrast to a composition profile of PFAs in water columns, PFOS
predominates in the tissues of all aquatic organisms, while in most cases PFOA
comprises only a minor portion of total PFAs analyzed (Table 4). Generally,
concentrations of PFOS in tissues increase with the trophic status of aquatic animals.
Thus, top predators in the aquatic food web, such as polar bears and piscivorous birds,
contain the greatest PFOS body burden up to low parts-per-million levels (Giesy and
Kannan, 2001). Together with the frequent prevalence of PFOA in the water column, the
bioavailability of PFOS seems to be greater than that of PFOA. A known precursor of
PFOS, PFOSA, was also ubiquitous in aquatic food web, and a large proportion of
PF OSA to total PFAs was observed in certain fish and invertebrates, indicating that these
organisms had limited capacities to biotransforrn PFOSA, possibly to the presence of
PFOS (Table 4). Longer chain PF CAs such as PFUnA and PFDA occur as the dominant
PF CAs in the tissues of aquatic organisms (Martin et al., 2004b). Increasing field-based
biomagnification factors (BMFs) and bioaccumulation factors (BAFs) with increasing
perfluoroalkyl chain-length were observed in laboratory bioaccumulation studies (Martin

et al., 2003a & 2003b; Martin et al., 2004b).

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Archived biological samples are useful to evaluate temporal trends of PFA
emission to the environment, assuming tissue data are a surrogate of PFA exposure.
PFOS concentrations in liver tissues of ringed seals (Phoca hispida) from east and west
Greenland showed a significantly increasing trend starting in 1986 (Bossi et al., 2005).
This annual increase was observed in PFDA and PFUA as well. Archived guillemot
(Uria aalge) eggs collected from 1968 to 2003 also revealed steady increases in PFA
concentrations by 7 to 11% annually (Holmstrdm et al., 2005). The results of lake trout
(Salvelinus namaycush) analysis were consistent with these observations in marine
mammals and birds, reporting a 4-fold increase between 1980 and 2001 (Martin et al.,
2004b). Future monitoring studies are warranted to see if early reduction efforts by
governments and relevant industries will result in a decline of PFA emission and

exposure.

Aquatic toxicology: Algae, invertebrates, amphibians, and fish

The adverse effects of water-bome PFAs, in particular PFOS or PFOA, have been
investigated in common test species (Table 5). Freshwater has been used most
commonly as the test medium, in part because the maximum test concentrations attained
in seawater are often lower than the expected effect concentrations. For example, the
solubility of PFOS in salt water is approximately 12.4 mg/L as compared to 680 mg/L in
freshwater. Toxicities of PFOA have been tested using different salts, mostly the

ammonium salt of PFOA (APFO).

23

 

 

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Literature values show that PFOS is moderately toxic under acute exposure
conditions and slightly toxic under chronic exposure conditions to aquatic plants,
invertebrates, and fish (Table 5.). In contrast, PFOA is rarely toxic to aquatic organisms
and adverse effects were only observed at concentrations of several hundred parts-per-
million (mg/L). Among commonly used test organisms, the aquatic insect Chironomus
tentans was the most sensitive species to PFOS exposure. Toxic effects such as mortality
and growth inhibition were observed at tens of part-per-billion levels (ug/L) in this
species, which were almost lOO-fold less than effect concentrations of PFOS in
freshwater fish (LCSO: 7.8-22mg/L). PFOS also exhibited comparable toxicity to marine
rainbow trout. Of the endpoints selected for fish species, fecundity was the most
sensitive. The 21-d 50% effect concentration (ECSO) on fecundity of fathead minnow
(Pimephales promelas) was determined to be 0.23 mg PFOS/L. The acute lethal
concentration which generates 50% mortality of an exposed population (LCSO) for PFOA
was quite comparable among fish species at a range of 569-740 mg/L, including rainbow
trout, fathead minnow, and bluegill sunfish. In amphibians, a partial life—cycle toxicity
test from early embryogenesis through complete metamorphosis was conducted using the
northern leopard frog (Rana pipiens). Time to metamorphosis was delayed and grth
was reduced in the 3 mg PFOS/L treatment.

A freshwater microcosm study was conducted for 42 d to evaluate the effects of
PFOS on zooplankton community structure and dynamics. A community-level no-
observable-effect concentration (NOEC) of 3.0 mg PFOS/L was determined for a 35-d
exposure. The most sensitive taxonomic groups, Cladocera and Copepoda, were virtually

eliminated at the greatest test concentration (30 mg PFOS/L) after 7 d. A threshold level

25

for the 35-d exposure was estimated to be between 0.3 and 3.0mg PFOS/L in this study.
Species richness was also negatively influenced. The total number of individual species
over the 35 (1 exposure in controls was 23.8, whereas only 8.2 species were found in the
30 mg PFOS/L treatment.

Biochemical and histological studies have been also conducted for PFAs using
various aquatic species. PFOS increased hepatic fatty acyl-CoA oxidase activity and
oxidative damage in fathead minnow after a 28-d PF OS exposure (Oakes et al., 2005). A
dose-dependent decrease in cell viability by PFOS and PFOA was found after a 24-h
exposure in primary cultured hepatocytes of freshwater tilapia (Oreochromis niloticus).
In that study, significant induction of reactive oxygen species was also found at an
exposure concentration of 15 mg PFOS/L and 15 mg PFOA/L, while antioxidant
enzymes activities and GSH content decreased. Though PFOS is a known peroxisome
proliferator in rat and mouse, the low sensitivity of fish species towards peroxisome
proliferators have been reported. For example, peroxisomal catalase and palmitoyl CoA
oxidase activities did not change significantly in PFOS-exposed carp (Cyprinus carpio)
(Hoff et al., 2003). A slightly increased incidence of thyroid follicular cell atrophy was
observed in metamorphs of the northern leopard frog exposed to 3 mg PFOS/L (Ankley
et al., 2004). Overall, it seems that acute or chronic effects are unlikely to occur at

reported environmentally relevant concentrations of PFOS and PF 0A in freshwater.

26

Study Design for Environmental Risk Characterization in Lake Shihwa

The Lake Shihwa area

Lake Shihwa, located on the west coast of Korea, is an artificial lake with a 12.7 km long
sea-dike (Figure 2). The lake receives both industrial and municipal wastewaters from the
neighboring Shihwa-Banwol Industrial Complexes (SBIC) and residential areas,
respectively. About seven thousands companies, largely steel/mechanical (48.9%),
electric/electronic (14.6%), and petrochemical companies (10.1%) currently operate in
SBIC. Initially, Lake Shihwa was constructed to supply freshwater for industrial and
agricultural purposes. Despite government efforts, gradual deterioration of lake water
quality has been reported as well as the massive death of certain bivalve species, due to
low annual precipitation and lack of enough tributaries in this region. For this reason,
pollution status in this area has been relatively well-documented compared to other
regions in Korea, and surveys indicated that lake seawater and sediments in upper inland
areas were contaminated with metals and trace organic contaminants (alkyl phenols,
PCBs and PAHs). Westerly winds dominate the lake at an annual average velocity of 1.5
m/s, and they typically flush the atmospheric contaminants emitted from SBIC to the
residential area on the east side of the lake. Since new management action allowing daily
exchange of lake waters with outer Kyounggi-bay seawaters in 2001 has been exercised,
water quality is reported to have improved compared to when the sea-dike was firmly
closed. In addition, a number of wildlife species, particularly birds, including migratory
species that use Lake Shihwa, are now observed in this region. Thus, it is timely and
imperative to determine the potential ecological risks of PFAs posed to Lake Shihwa

wildlife.

27

 

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et al., 2006).

28

 

PFA Concentrations in Lake Shihwa Waters

Only a small number of publications are available for PFAs pollution in Korean waters.
The first peer-reviewed report dates back to 2004 as a part of screening study to survey
PFA contamination in coastal waters of highly industrialized East Asian regions,
including Hong Kong, South China, and Korea. Surprisingly greatest concentrations of
PFAs were measured at some locations in Korea. One sampling location situated within
Kyeonggi Bay had 0.33 ng PFOSA/L, 730 ng PFOS/L, 13 ng PFNA/L, 320 ng PFOA/L,
and 52 ng PFHS/L. These concentrations were 5 to 63-fold greater than the maximum
concentrations determined in other locations in that international joint study. Except for
this hotspot, PFA concentrations in Korean coastal waters were comparable to other East
Asian locations.

This survey result raised the urgent question of the sources and contamination
status in Kyeonggi Bay and Lake Shihwa, both of which are heavily influenced by
industrial effluents from SBIC. As a result, the Lake Shihwa region was revisited in
December 2004 and a more systematic monitoring plan was made to ensure the quality of
data, including sampling locations, QA/QC for chemical analysis, and the latest
instrumentation. Water samples were collected from streams discharging industrial
effluents into Lake Shihwa (n=21), from Lake Shihwa (n=8), and Gyeonggi Bay (n=5).
In that study, a broad range of PFAs were monitored, including four sulfonic acids (PFOS,
PFHxS, PFBS, and FOSA) and five carboxylic acids (PFDA, PFNA, PFOA, PFHpA, and
PFHxA). Concentrations of all target PFAs were greater in inland streams than in the
lake itself or in Kyeonggi Bay waters (Table 6.). Of the PFAs investigated, PFOS and

PFOA were detected at all sampling sites and occurred at the greatest concentrations.

29

Concentrations of PF OS were in the range of 2-651 ng/L, while PFOA was present at 10-
fold lower concentrations with a range of 09-62 ng/L. An interesting finding was a
decreasing concentration gradient of PFAs as a function of distance from inland waters,
implying that industrial effluents were a possible release source into Lake Shihwa and

Kyeonggi Bay.

30

 

 

 

 

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31

Hazard Assessment Strategy

Intensive monitoring efforts for PFAs around the globe have generated a lot of useful
data, which can be used by regulatory agencies to take the proper management actions to
prevent or reduce the risks of PFA exposure to humans and wildlife. However, such
efforts are rarely found in Korea, one of the most highly industrialized countries in the
world. Despite the recent study on PFAs in lake waters, contamination status of these
emerging chemicals in living animals are not understood at all. This study was designed
to conduct a preliminary risk assessment of PFA exposure to marine species and their top
predators (birds) in the Lake Shihwa region, where one of the more industrialized
complexes in Korea is located. In order to achieve this goal, a research framework was
developed (Figure 3).

In the first chapter, the peer-reviewed literature on PFAS was studied from
toxicology to chemistry and regulation. The available toxicological benchmark values
protective of fish and avian species were also determined. In Chapter II, the
pharmacokinetics of PFAs were investigated using implanted white leghom chickens
(Gallus gallus) as model species under laboratory conditions for an exposure and
elimination period of 8 weeks in order to assess the risk of PFAs to bird species. This
exposure study was collaborated with the National Institute of Animal Health, Tsukuba,
Japan. Briefly, capsules containing lPFOS and PFOA were subcutaneously implanted in
an experimental animal for 4 weeks and they were allowed to depurate for a further 4
weeks. Blood samples were taken at an interval of two or three days over the entire

experimental period. Organ samples such as brain, kidney, and liver were collected at the

32

termination of the exposure and after depuration, respectively. Uptake and depuration

kinetics were studied separately in blood and organs and the results were discussed.

Chapter 1

Chapter 2

Chapter 3

Chapter 4

A
‘

%

Overview of dissertation research
- Literature review
- Research planning

Phannacokinetlc Study

- Experimental chicken

- PFOA and PFOS implantation
- Sampling of blood and organs
- 4 wk exposure period

- 4 week elimination period

Instrumental Analysis - LCIMSMS

- Training and qualification for use

- Visit to AIST, Japan and SUNY at Albany
- Analytical method development

- Sample analysis and QA/QC

- Data interpretation

Exposure Assessment - Fieldwork
- Field trip to Lake Shihwa, Korea
- Preliminary survey for sampling area
- Aquatic organisms sampling
(fishes, mussels, oysters, and blue crabs)
- Bird eggs sampling

Toxicological Study

- In vitm cell bioassay training

- Gap junction cellular communication

- Derivation of relative potency of PFA

- Calculation of PFOS-equivalent concentration

Conclusion
- Summary of research

Figure 3. Dissertation research scheme and assigned works in each chapter

33

In Chapters III and IV, PFA concentrations were determined and associated risks

were characterized in marine organisms such as fish, mussel, oyster, and blue crab, and in
egg yolks of birds, little egret (Egretta garzetta), little ringed plover (Charadrius dubius),
and parrot bill (Paradoxomis webbiana). A preliminary field survey and field sampling
were conducted from May to June 2006 in the Lake Shihwa area. Stationary fish nets
were installed for 4 days to collect marine species. Afier a preliminary survey for nests
with a help of an avian expert, birds’ eggs were sampled. Nests of little ringed plover
were found at islands in Lake Shihwa and eggs of parrot bill were collected in the
wetland of Lake Shihwa away from known pollution sources. One large little egret
colony was surveyed in a section of the city of Ansan, Korea.
In biota, PFAs were present as a mixture of PFAs rather than a single congener. As
discussed earlier, unfortunately most toxicological studies targeted PFOS and PFOA.
Currently, water quality guideline values that are protective of aquatic organisms are only
available for PFOS. Recently, toxicity reference values (TRVs) of PFOS for blood, liver,
and egg yolk were developed for avian species. Thus, the risk assessment approach used
for exposure to complex mixtures of PCBs was benchmarked for a mixture of PFA in
biotic samples.

Because PFAs are the newest halogenated contaminants, the use of the proper
analytical techniques as well as the latest instruments are essential to the success of the
present research. To analyze the biological samples from the pharmacokinetic study and
fieldwork campaign, I visited the National Institute of Advanced Industrial Science and
Technology (AIST), Tsukuba, Japan for a total of nine months. The potential pollutants

group at AIST is one of the renowned laboratories for the analysis of PFAs. They

34

developed high-resolution liquid-chromatography (HPLC) tandem mass-spectrometry
(LC/MSMS) solely devoted to PFAs analysis. For some remaining samples, I also
visited the State University of New York at Albany, NY, and completed sample analysis

for a month with a help of Prof. Kannan.

35

Chapter 2
Depuration Kinetics and Tissue Disposition of PFOA and PFOS in White Leghorn

Chickens (Gallus gallus) Administered by Subcutaneous Implantation

Abstract

To address the elimination kinetics and tissue disposition of perfluorooctanoic acid
(PFOA) and perfluorooctanesulfonate (PFOS) in avian species, an exposure study was
designed with low-exposure conditions using a subcutaneous implantation in male
chickens (Gallus gallus). Chickens were exposed to two levels of PFOA or PFOS for 4
weeks and allowed to depurate for additional 4 weeks. Over the entire experimental
periods, body index, clinical biochemistry and histopathological assessments were not
statistically different among treatments relative to the controls (p>0.05), except
significant decreases in total cholesterol and phospholipids in PFOS-exposed chickens.
The elimination rates for PFOA (0.150i0.010 d") were almost six-fold faster than those
of PFOS (0.0233c0.004 (1"). After an exposure, kidney had the highest concentration of
PFOA followed by liver and brain. In contrast, liver was found to be a central
accumulation site for PFOS with lesser concentrations in kidney and brain. The
estimated biological half-life of PFOA (4.6:t0.3 d) and PFOS (ll/2 = 125 d) in chickens

was simlar to mammals.

36

Introduction

The use of perfluorinated acids (PFAs) in a variety of products has resulted in them being
ubiquitous in the environment, occurring globally in humans and wildlife (Giesy and
Kannan, 2002; Olsen et al., 1999; Lau et al., 2004). PFAs have been used as surface
protectors for carpets and leather, and as surfactants in cosmetics as well as processing
aids in the production of fluorinated polymers and active-component in fire-fighting
foams (Kissa, 2001). To date, due to their widespread use in industrial and commercial
applications the two PFAs that have received the greatest attention are perfluorooctanoic
acid (PFOA) and perfluorooctane sulfonate (PFOS). In a monitoring study of human
serum collected from the general public, PFOS was determined to be the most abundant
PFA followed by PFOA (Kannan et al., 2004). Similar profiles of PFAs have also been
observed in the tissues of terrestrial and marine wildlife with PFOS being the dominant
PFA in these samples with concentrations in some cases exceeding those measured in
human populations (Houde et al., 2006). Several food web studies have shown that some
PFAs have the potential to bioaccumulate into the lower trophic organisms and through
tropic transfer and biomagnification, can accumulate in to upper trophic level organisms
(Vijver et al., 2003; Martin et al., 2004). While exposure pathways of PFOA, PFOS and
related PFAs to humans have not yet been fully elucidated, the consumption of fish
(Falandysz et al., 2006) and farm animals (Guruge et al., 2005) have been suggested as
major contributors of PFAs to exposed human populations. Nevertheless, efforts
reducing PFAs environmental emission are now being initiated in both governmental and

industrial areas.

37

To effectively control exposures to these compounds, it is necessary to understand
their pathways of exposure and to develop models to predict the rates of movement in the
environment. However, due to their amphiphilic properties, PF As do not behave in the
same manner as the more studied organochlorine contaminants. Specifically, PFAs have
fewer tendencies to partition into lipids, but rather are preferentially bound to proteins
and retained in the blood and liver of wild animals. To date, only few pharmacokinetic
studies with PFOA and PFOS have been conducted with mammals that have included
rats, dogs, and monkeys (Seacat et al., 2002; Kudo et al., 2002; Lau et al., 2004). In
these studies, administered PFOA has shown to be readily absorbed but has a relatively
short half-life with notable species- and gender- differences in rates of elimination when
compared to PFOS. For example, in male rats the half-life for PFOA of blood is
approximately one week while in female rats the half-life is approximately one day
(Butenhoff et al., 2004). In contrast, PFOS is also readily absorbed but is poorly
eliminated from blood and other tissues with biological half-lives that can range from
several weeks to months depending on species and sex.

Concentrations of PFAs in wild birds have been studied globally, and data suggest
that birds from urban areas are more contaminated with PFAs than those from rural areas
(Kannan et al., 2001; Verreault et al., 2005). Other studies have quantified PFOS in
several farm animals including chickens and livestock. In that study, concentrations of
PF OS in blood plasma and liver were greater in chickens than in other farm animals that
were evaluated (Guruge et al., 2005). While pharmacokinetic studies with PFOS have
not been conducted with avian species, some kinetic data are available from acute and

chronic dietary studies that been conducted with two species, northern bobwhite quail

38

(Colinus virginianus) and the mallard (Anas platyrhynchus) (Newsted et al., 2005). The
results of those studies indicated that the half-life of PFOS in the blood and liver of
juvenile birds ranged from approximately 7 to 18 d while in adult birds, the half-life
blood ranged from 14 to 21 d. Based on the results of those avian studies, the elimination
rate of PFOS from birds has been assumed to be faster than those observed in mammalian
species. However, since treatment levels used in these studies were greater than that
usually used in kinetic studies, the overall pharmacokinetics of PFOS in those species
may have been influenced. Therefore, additional kinetic studies with PFOA and PFOS
for birds are needed under low-exposure levels that would not be expected to affect
pharmacokinetic parameters.

To address the above questions, an exposure study was designed with low-
exposure conditions to better understand the pharmacokinetic behavior of PFOA and
PFOS in male chickens (Gallus gallus). First, elimination kinetic parameters from
chicken blood were measured for PF OA and PFOS administered subcutaneously. Second,
tissue disposition patterns of introduced PFOA and PFOS were examined in tissues of
brain, kidney, and liver following an exposure and a elimination period. In addition to
pharmacokinetic evaluations, biochemical and histopathological parameters were also
evaluated. Finally, the pharmacokinetic results from this study were compared with
findings from other studies that have been conducted with alternative study designs
and/or other species. A subcutaneous implant-exposure scheme, which is widely
exercised in veterinary science as an efficient drug delivery system, was used to introduce
the PFOA or PFOS into the chickens. For this reason, the uptake rate kinetics were not

be determined in this study. To date, chickens have been shown to be among the most

39

sensitive avian species to PFOS and, accordingly these data will aid in future ecological

risk evaluations of PFAs exposure for avian species (Molina et al., 2006).

Materials and methods

Test substances and reagents

Two perfluorinated chemicals, perfluorooctanoate (PF OA, purity 95%, CAS number 33 5-
67-1) and perfluorooctane sulfonate (PFOSK, purity >98%, CAS number 2795-39-3)
were purchased from Wako Chemicals, Japan and Fluka, Italy, respectively. Pesticide-
grade methanol, ammonium acetate, and ammonium solution (25%) were purchased from
Wako Chemicals, Japan. Milli-Q water was used in the whole experiment. Nylon filters
(0.1 um, 13 mm id) were purchased from Iwaki, Japan. Additional clean-up for tissue
extracts was carried out using Oasis weak-anion exchange (WAX®) cartridges purchased

from Waters Corp., Milford, MA, USA.

Animals and Exposure

This experiment followed the guidelines for animal experiments of the National Institute
of Animal Health, Tsukuba, Japan. White leghom (G. gallus) PDL-l strain were
obtained from a flock for which performance of 3 successive generations of specified-
pathogen-free chickens had been maintained to insure the health of the birds used in the
study. Eggs were hatched and male chickens were housed until six-weeks of age at
temperature and humidity-controlled facilities located at the National Institute of Animal
Health, Japan. Hatchlings were fed a standard experimental diet (SDL-l) while chickens

greater than 4 wk of age, were fed SDL-4 O‘lippon Formula Feed Co., Ltd). Six-week-

40

old, male chickens (347.4 :l: 15.7 g, n = 30) were randomly selected and placed into cages,
one cage per treatment with six chickens per cage. Experimental treatments consisted of
chickens exposed to either PFOA (0.1 mg/mL or 0.5 mg/mL) or PFOS (0.02 mg/mL or
0.1 mg/mL) or a saline vehicle control. All stock solutions were prepared in 0.9% NaCl
in Milli-Q water. Exposure of the chickens to these concentrations was done via the
subcutaneous implantation of a 2 mL osmotic pump (ALZET® 2ML4), which has a
releasing rate of 2.5 uL/hr for 4 wk. Under sodium pentobarbital anesthesia, an osmotic
pump was implanted surgically into hypodermal tissue at right side of trunk of each
chicken. Chickens were fed with the standard SDL—4 diet during the exposure and the
elimination phase of the study. At intervals of 2 to 3 d, one to two milliliter of blood was
drawn from the wing vein of chickens during experimental periods using a heparinated
needle and stored in polypropylene (PP) centrifuge tubes until analyzed for PFOA or
PFOS. To avoid the potential influence from the hypodermal pump operation, blood was
sampled from the wing vein opposite to the wing where the pump had been implanted.
At the end of an initial 4 wk-exposure period, half of the chickens from each treatment
group were anesthetized and blood was collected for determination of blood chemistries.
Following a blood collection, the organs of the euthanized chickens, such as brain, liver,
and kidney were placed in PP bags for tissue analysis and histopathology examination.
The remaining chickens were maintained for an additional 4 weeks and then euthanized
when blood and other tissue were then processed as given above. In an elimination phase,
implanted capsules were not retrieved from experimental chickens in order to avoid

additional surgery; therefore it was assumed that all perfluorinated chemicals in implants

41

had been in the proceeding 4-wk period. All biological samples (blood and organs) were

kept at negative 20 C° until instrumental analysis.

Sample extraction

Blood samples were extracted with an ion-pairing method with some modifications
(Kannan et al., 2004). Briefly, 0.5 mL of blood sample was diluted 10 times with saline
buffer (0.9% NaCl in Milli-Q water) while organ tissues were mechanically homogenized
with a vortex mixer and 0.5g of the homogenate was diluted with 2 mL Milli-Q water.
One milliliter of diluted blood or tissue-water mixture was then transferred into a 15 mL
PP tube, and 1 mL of 0.5 M ion-pairing agent (Tetrabutyl ammonium adjusted to pH 10)
was added to the mixture. Two milliliter of 0.25 M extraction buffer (Sodium carbonate
+ Sodium bicarbonate) was then added followed by the addition of 5 mL of methyl tert-
butyl ether (MTBE). Samples were shaken for 20 min then centrifuged for 15 min (2000
rpm). The organic phase was removed (4 mL) and put in a clean 15 mL PP centrifuge
tube. The extraction was then repeated twice and the organic phase from all extractions
were combined and then evaporated to near dryness under a gentle stream of nitrogen.
The sample was then re-dissolved in 1 ml methanol and then filtered through 0.1 um
nylon filter. For the tissue analysis, a solid-phase extraction (SPE) step was employed as
an additional clean-up with some modifications (So et al., 2006). Briefly, a half milliliter
aliquant of unfiltered organ extract obtained from ion-pairing extraction was diluted with
100 mL Milli-Q water and then the water-extract mixture was passed through an Oasis
WAX® cartridge (0.2 g, 6 cm3) at an elution rate of l drop/sec. At the completion of

sample loading, the cartridge was washed with buffer adjusted to pH 4 (25 mM acetic

42

acid 170 mL + 25 mM ammonium acetate 30 mL) and methanol. Then, a target fraction
containing PFOA or PFOS was collected with 0.1% NH40H dissolved in methanol.
Teflon or glassware was avoided in extraction procedures to remove possible

contamination of samples and sorption of analytes.

Matrix recoveries

To evaluate overall extraction efficiencies, either PFOA or PF OS was fortified into blood
and organs tissue homogenates prior to extraction. Recoveries from blood (n=10) were
81.3:t6.7% and 87.0:l:5.3% for PFOA and PFOS, respectively. In tissues, comparable
efficiencies were obtained in brain, kidney, and liver. Recoveries from PFOA-spiked
samples ranged from 86.9 to 94.3% (n=4), while extractions efficiency ranged from 81.4
to 88.2% (n=4) for PFOS-spiked tissues. Reported concentrations of PFOA and PFOS
were not corrected for recoveries of matrix spikes. The limit of quantification (LOQ) for

both compounds varied from 1 to 5 ng/mL or ng/ g wet wt., depending on the sample type.

Instrumental analysis and data analysis

Quantification of PFOA and PFOS in blood or tissues was conducted using HPLC with
high resolution, electrospray tandem mass spectrometry (HPLC-MS/MS). Separation of
analytes was performed by an Agilent HP 1100 liquid chromatography (Agilent, Palo
Alto, California) interfaced with a Micromass Quattro II mass spectrometer (Waters
Corp., Milford, Massachusetts) operated in electrospray negative mode. Ten uL aliquot

of extract was injected onto a Keystone Betasil Clg column (2.1 mm id. X 50 mm length,

5 um) with 2 mM ammonium acetate and methanol as mobile phase, starting at 10%

43

methanol. Ions were monitored using selected reaction monitoring at m/z 413 and 369
for PFOA and at m/z 499 and 99 for PFOS. Concentration of PFOA or PFOS in extracts
was quantified using calibration curves constructed by external standards (0.01, 0.05, 0.2,
1, 10 ng/mL). Acquired data were deemed acceptable if QC standard included in sample
batch fell within 30% of the theoretical value, otherwise samples were run again with a

new calibration curve.

Clinical chemistry and histopathology

Thirteen biochemical parameters were analyzed in plasma using a Hitachi 7020 auto-
analyzer with standards from Wako Pure Chemical Industries Ltd, Japan. All standards
were used in accordance with the manufacturer’s instruction and stated expiration date.
Parameters included total cholesterol (T-Cho), free total cholesterol (F-Cho), high-
density lipoprotein (HDL), low-density lipoprotein (LDL), total protein (TP), albumin
(Alb), blood urea nitrogen (BUN), non-esterified fatty acids (NEFA), , phospholipids
(PL), aspartate aminotransferase (AST), alanine aminotransferase (ALT), alkaline
phosphatase (ALP), lactate dehydrogenase (LDH). The following tissues were fixed in
10% phosphate—buffered formalin and processed for histopathological examination: liver,

kidney, spleen, heart, lung, thymus, testis, bursa of fabricius, and brain.

Data and statistical analysis
A one-compartment model was used to describe the elimination behavior of PFOA or
PFOS from blood in male chickens (Eq. 3).

Ct = Coexp(-kt) (Eq. 3)

44

Where Ct is the concentration of PFOA or PF OS at the time (t) in an elimination phase,
Co is the concentration at the onset of depuration (ng/mL).

To account for growth dilution as a factor in determining the elimination rate
kinetics, elimination rate constant was determined by first estimating the overall
elimination rate constant then adjusting the rate constant by the growth rate constant for
each chicken used in the study (Eq. 2).

k = k’ + k, (EQ- 2)
Where k is the overall elimination rate constant (day"), k’ and kg are final first order
elimination rate constant (day") and the grth rate constant (day'l), respectively.

To evaluate treatment effects on body index, serum chemistry, and tissue
accumulation, one-way ANOVA was performed with SYSTAT® at the significance level

set to p=0.05.

Results

Body index, clinical biochemistry and histopathology

No significant differences were observed for body-weight gains among doses (vehicle
control, low-dosed and high-dosed) for PFOA and PFOS nor were there any statistical
differences observed between PFOA and PFOS treatments over an entire experimental
period (p > 0.05) (Table 7). Growth rates (20 ~ 21 g/d) were determined to be
comparable between non-exposed and exposed chickens at the end of a 4—wk exposure
phase. Following an exposure period, there were slight increases in the liver to body
weight ratios from PFOA and PFOS treated groups (2.5 ~ 2.6), however these increases

were not statistically different with the vehicle control group (2.2). This result was also

45

observed in chickens collected at the end of a depuration period of the study. Exposure
to PFOA or PFOS also did not statistically affect either the kidney to body ratio or the
testis to body ratios. Sampled amount of brain were so small that we could not use it for
comparison. In chickens collected at the termination of an exposure period, most of
clinical chemistry parameters were not significant different among treatments (Table 8)
and no significant lesions were seen relative to those from vehicle controls (data not
shown). However, after a depuration phase there were significant decreases in total
cholesterol and phospholipids in chickens exposed to both low- (0.02 mg PFOS/mL) and

high-dosed (0.1 mg PFOS/mL) treatment.

46

Table 7. Mean body-weight gain (s.d.) and organ to body weight ratio at the end of
an exposure (A) and an elimination period 03).

 

 

 

 

 

A. Exposure Body wt. (g) Liver/Body (%) Kidney/Body (%) Testis/Body (%)
Vehicle control 465 ( 4) 2.24 (0.19) 0.96 (0.01) 0.039 (0.001)
Low PFOA 468 (38) 2.63 (0.23) 1.01 (0.09) 0.043 (0.012)
High PFOA 477 (18) 2.48 (0.12) 0.97 (0.05) 0.034 (0.005)
Low PFOS 495 (32) 2.51 (0.24) 1.03 (0.03) 0.045 (0.025)
High PFOS 505 (93) 2.51 (0.21) 0.95 (0.10) 0.047 (0.022)
B. Elimination Body wt. (g) Liver/Body (%) Kidney/Body (%) Testis/Body (%)
Vehicle control 1063 ( 4) 2.07 (0.14) 0.83 (0.10) 0.12 (0.11)
Low PFOA 1000 (100) 1.94 (0.21) 0.84 (0.01) 0.27 (0.14)
High PFOA 965 ( 71) 1.94 (0.01) 0.85(0.01) 0.10 (0.02)
Low PFOS 1038 ( 28) 1.97 (0.12) 0.87 (0.12) 0.19 (0.07)
High PFOS 1015 (135) 1.82 (0.10) 0.81 (0.05) 0.22 (0.16)

 

Note. Six week-old male chicks were exposed to either PFOA (total 0.2mg or 1.0mg) or
PFOS (total 0.04mg or 0.2mg) via subcutaneous implantation for 4 weeks and were

allowed to depurate for further 4 weeks (n=6 for exposure and n=3 for depuration).

47

 

 

 

 

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Uptake profiles and elimination behaviors in blood
The time-course concentrations of PFOA and PFOS released from implanted capsules
demonstrated different uptake profiles in chicken blood over an exposure period.
Introduced PFOA in both low- (0.1 ‘mg/mL) and high-dosed (0.5 mg/mL) treatment
behaved similarly with a rapid increase (tmax=7 d) followed by sustained blood levels in
later periods, while blood-bome PFOS fluctuated over a 4wk-uptake period (Figure 4).
Concentrations of PFOA and PFOS in blood reflected the dose from capsules containing
target PFAs. For example, the maximum PFOA concentrations in blood taken from low-
dosed and high-dosed treatment were 103 and 333 ng/mL, respectively. In non-exposed
chickens during an uptake period, PFOA blood concentrations were below the limit of
quantification, while measured PFOS concentrations (3.5i0.7 ng/mL) were at least 8
times lower than those in any given PFOS treatment.

Rates of elimination from blood in chickens were significantly different for PF OA
and PFOS (Figure 5). Concentrations of PFOA in blood decreased rapidly during a
depuration phase, with PF OA concentrations at the termination of depuration only being
2~3% compared to those in the onset of a deputation. In contrast, only 48~52%
reduction of PFOS concentrations in the blood compartment was found during the same
experimental period (Figure 5). The first-order elimination rate constant was similar
between treatments administered with the same compound (PFOA: 0.150i0.010 (1" and
PFOS: 0.023i0.004 d'l). The overall elimination half-life (ll/2) of treatment-averaged
PFOA (4.6 d) was almost six-fold shorter than that of treatment-averaged PFOS (31.1 (1).
Growth dilutions that occurred in a depuration period did not influence on the final t1 )2 of

PFOA (5.2 d) on average. However, it increased the elimination half-live of PFOS such

49

that high-dosed treatment (ti/2 = 72.2 (1) had smaller value than low-dosed treatment (ti/2
= 177.7d). In a vehicle control group, PFOA and PFOS blood concentrations over a

depuration period were comparable to those concentrations in an exposure period.

50

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52

Organ concentrations

Distinctive accumulation features of these two perfluorinated compounds were
observed in organs (brain, kidney, and liver) of experimental chickens following an
exposure and a depuration period (Figure 6). After a 4-wk dosing period, the greatest
concentration of PFOA was measured in kidney tissue, followed by liver and brain tissue
(Fig. 6A~C). In the high-dosed treatment (total 1.0mg PFOA dosed), the concentration
of PFOA in kidney was as high as 186 i 40 ng/g w. A measurable concentration of
PF 0A was also detected in brain homogenate of PFOA-treated chickens. Concentrations
of PFOA in all tissues rapidly decreased proportionately to that in blood. By the
completion of a depuration phase, an average of approximately 92% of organ PFOA was
cleared from all tissues investigated. Organs of unexposed chickens did not have any
contamination of PFOA in entire experimental periods. Interestingly, tissue disposition
and elimination pattern of PFOS were significantly different than those observed for
PFOA. Following a cessation of exposure, the greatest concentration of PFOS was
determined in liver tissues, with lesser concentrations in kidney and brain (Fig. 6D~F).
Hepatic samples of chicken from the high-dosed treatment (total 0.2mg PFOS dosed)
accumulated up to 769 ng/g ww, which was 5- and 55-fold greater than those in kidney
and brain, respectively. Even after a 4-wk depuration period, concentrations of PFOS in
organs were not significantly changed, while PFOS level increased in brain tissue of the
least dosed chickens (p < 0.05). Experimental chickens incubated with only saline buffer
also contained detectable amounts of PF OS in liver tissues in all experimental periods,

and brain and kidney tissues only after a depuration phase.

53

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54

Discussion

PF 0A vs. PF 0S elimination kinetics

The observation that introduced PF OS was retained longer in the G. gallus than PF 0A is
consistent with the empirical results from other animals (Butenhoff, et al., 2004; Lau et
al., 2004). This different elimination kinetics is interesting, in that these two compounds
are structurally analogous anionic surfactants and possess comparable physico-chemical
properties such as a low pKa (presumably <3), a low Henry’s Law constant, and similar
hydrophobicity based on critical micelle concentration (CMC for C7F.5COOH = 8.7 ~
10.5 mmol/L; CMC for C3F17SO3K = 8 mmol/L) (Kissa, 2001). In contrast, they can be
distinguished with their hydrophilic carboxylic and sulfonic functional group that endow
them with surfactant properties and the fact that overall, PFOS contains one additional
carbon in the perfluorinated portion of the molecule.

To date there has been no clear mechanistic explanation of the dissimilar
elimination behaviors among PFAs with few suggested mechanisms. At present, urinary
excretion is known to a major pathway for clearing PFOA from mammals, most likely by
the process of renal tubular secretion (Harada et al., 2005). Following an intra-peritoneal
administration, 55% of closed PFOA was eliminated in both male and female rates within
120 h after an injection, while only less than 1% was eliminated in the feces (Kudo et al.,
2001). The level of organic anion transporters (OATS) mRNA present in kidney of the
male rat was only 13% of that present in the female rat whose clearance rate is greater
than the male rat (Kudo et al., 2002). Surprisingly, castrated male rats showed the
increased expression of OAT2 mRNA and competent eliminating ability of PFOA to that

of female rats, which suggests an important role of these transporter proteins for PFOA

55

removal via the kidney and also influences of sex-hormones (Kudo et al., 2001; Kudo et
al., 2002). In a case of PFOS, removal mechanisms such as an active transport via OATs
have not been well examined. Rather than facilitating of PFOS removal, retardation
processes of PF 08 removal could be hypothesized such as a well-known enterohepatic-
recycling. In addition, greater binding strength of PFOS for serum proteins may also
contribute the increased retention of PFOS than PFOA, by reducing renal excretion
efficiency by mitigating water-soluble nature of its conjugate base (C3F17SO'3) dominant
species at physiological pH. For example, displacement potency of PFOS for native
hormones in chicken serum was greater than that of PF OA (Jones et al., 2003). It may be
useful to study comparatively the modulation of expression of OATS or other transporters
in kidney after a PFOS exposure, which would help to elucidate the observed differences

between the clearance rates of PFOA and PF OS in birds.

Organ distribution

Tissue disposition of these persistent and anionic PFAs are different from that of
conventional persistent, but nonionic organic contaminants such as PCBs and PAHs,
which tend to accumulate in fat tissue. In contrast, the primary sites of accumulation for
PFAs are found to be in blood, liver, and kidney depending on which species are
investigated. One of the reasons for the favored accumulation in those sites is the binding
of PFAs to protein albumin, which is abundant in blood and liver (Jones et al., 2003;
Martine et al., 2003a). Though the octanol-water partitioning coefficients (Kow) are
incorporated in a bioaccumulation modeling for hydrophobic compounds, the

amphiphilic nature of PFCs makes calculation of K0W impossible. Even though a K0w for

56

PFOA or PFOS could be driven using its solubility in water and octanol determined
separately, an estimate predicts less potential to accumulate in adipose tissue.

Among the organs and blood, kidney accumulated the greatest concentrations of
PFOA with its relative concentration being intensified toward the end of a depuration
period (Figure 7). As already discussed, renal uptake processes by the OAT family for
removal of PFOA may also contribute accumulation of PFOA in kidney tissues as an
intermediate step before urinary excretion. Efficient exclusion of PFOA organ residue
once an exposure source was removed may also imply the dynamic redistribution of
organ-accumulated PFOA into the blood compartment where subsequent active
elimination processes follow. Liver is recognized as a central site for PFOS storage, thus
this tissue has been frequently analyzed in monitoring studies of PFAs (Houde et al.,
2006). Recirculation of PFOS via enterohepatic system has been suggested as one reason
for extended retention of PFOS in the body (Johnson et al., 1984; Lau et al., 2004). In
aquatic exposure studies, it was shown that relatively great amounts of PFOS were
accumulated in organs related to enterohepatic cycling such as the gall bladder and liver
(Martin et al., 2003a). Dietary PFOS was more efficiently absorbed into trout intestine
than food-bome PFOA, supporting the enterohepatic recirculation hypothesis (Martin et
al., 2003b). In all tissues examined, PFOS concentrations were not significantly varied
between a cessation of exposure and depuration, rather a statistically significant increase
of PFOS level was observed in brain tissue. The ratio of PFOS concentration in liver to
that in blood from low-dosed treatment (PFOSHVCEbIood = 15) was three times greater than
that determined in high-dosed treatment (PFOSnmblood = 5) following an exposure period,

but was similar at the end of a depuration phase with increased ratios (PFOSnvcr;blood = 33

57

and 29 for low- and high-dose treatment, respectively). This implies that either the
mechanism of elimination of PFOS from liver is a slow process compared to that from
blood, or blood-bome PFOS may be redistributed to liver. This liver to blood ratios
could be different in wildlife, where'precursors of PFOS either identified or not also
contribute overall bioaccumulation of PFOS which is the ultimate degradation product of

perfluorooctane sulfonyl fluoride based products (Giesy and Kannan, 2002).

58

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60

Analysis of the relative mass of PFOA or PFOS in each organ following an
exposure and a depuration phase demonstrated that blood and liver are the largest
reservoirs controlling PFOA and PFOS body-burden in exposed chickens, respectively
(Figure 8). The estimated total quantity of PFOA or PFOS that remained in bodies (=
blood + brain + kidney + liver) of chickens at the end of a depuration period was only
0.7% and 20.4% of total PFOA and PFOS administered via subcutaneous implantations.
This small PFOA body residue is in part attributed to rapid removal characteristics of
PFOA from the body. However, this mass balance of introduced PFOA and PFOS has
some uncertainty. First of all, we did not retrieve the introduced capsule containing
PFOA or PFOS for residual determination at the completion of treatments. We did not
measure PFAs concentrations in chicken flesh that accounts for the most of body weight
due to the lack of proper extraction method.

Age—dependent organ accumulation of PFOS was observed in non-exposed
chickens in this study (Figure 6D-F). Until the first 4-wk exposure phase was over, all
the organ PFOS concentrations were not measurable; however significant PFOS levels
were measured in brain, kidney, and liver tissue following a 4-wk extended incubation.
The effects of age on PFOS accumulation were observed in studies with domestic farm
animals (Guruge et al., 2005). In general, PFOS concentrations in liver and blood
increased with ages of farm animals investigated. PFOS concentration in the diet of
experimental chickens was not determined in this study. However, the influences of
supplementary food as a likely exposure pathway to livestock need to be further
evaluated, since consumption of domestic animals could be an exposure route of PFAs to

human population. Detection of PFOA and PFOS in brain of the exposed chickens

6l

suggests that in birds which are exposed to PFAs, these anionic compounds may cross the
blood-brain barrier that is vital to inhibit entry of xenobiotic contaminants into the central
nervous system. Similar to this laboratory scale exposure study, measurable
concentrations of PFOS were reported in brain tissues of wild birds such as glaucous
gulls and pelicans (Verreault et al., 2005; Olivero-Verbel et al., 2006). However, careful
interpretation should be taken with brain data, because anatomical complexity of brain
introduces PFOS contamination through extraction errors such that blood capillaries are
mistakenly included in tissue homogenate. In addition, the fact that only the free anion
can travel into brain mitigates the possibility of PFOS or PFOA passage, presumably

bound to plasma proteins (Kaassen et al., 2001).

Sublethal eflects

The observation of statistically significant lesser total cholesterol and
phospholipid levels in plasma after the PFOS exposure is consistent with other PFOS
exposure studies (Seacat et al., 2003). Inhibition of the rate-limiting enzyme, HMG-CoA
reductase in cholesterol synthesis was observed in PFOS-exposed rats (Seacat et al.,
2003). Interestingly, the significant decline of total cholesterol and phospholipids at the
end of a depuration phase rather than at the end of an exposure phase suggests the current
exposure levels and/or 4-wk exposure duration were not enough to cause altered lipid
metabolism in chickens, but prolonged PFOS exposure until the depuration phase
pronounced the reduction of those lipids levels. However, these changes did not affect
histopathology of the exposed chickens with any significant lesions being not seen in

examined organs of chickens from treated groups. It has been suggested that high PFOA

62

exposure could interfere with genes involved in fatty acid metabolism and cholesterol
syntheses in rats (Guruge et al., 2006), however fatty acid related parameters were not

changed in any PFOA treatments in this experiment.

Comparison with other studies

The biological half-life of PFOA and PFOS in G. gallus were compared with
other male species and summarized in Table 9. The half—life of PFOA from the chicken
(ti/2 = 5.2 d) were comparable to that of the rat (ti/2 = 5.6 d), and about 4 times faster than
that of the dog (m2 = 20 ~ 23 d) and monkey (t1 ,2 = 21 (1). Thus, the half-life seems to be
a fimction of body size and/or physiological parameters, such that fish (tug = 4.5 d) have
the fastest clearance of PFOA, followed by the chicken=rat>dog=monkey> human
(Harada et al., 2005). In contrast, the biological half-life of PFOS in our treated chicken
(t1 /2 = 125 d) was similar to mammals including the rat (t1 /2 > 90 d) and the monkey (t1 ,2 =
100 ~ 200 d). This suggests, unlike PFOA body behavior, that a number of vertebrates
share common accumulation mechanisms for PFOS, through binding to plasma proteins
as a governing process that is unlike in invertebrates (e.g., fish PFOS t1 ,2 = 12 d).

The only available pharmacokinetic study with two bird species reported a shorter
biological half-life for PFOS than those obtained in our experiment (Newsted et al.,
2005). Several factors may have influenced the pharmacokinetics in the two
experimental animal studies. First, in our study we used a subcutaneous implantation
method, while they employed a dietary exposure dosing (Newsted et al., 2005).
Metabolism of target compound introduced often affects pharmacokinetic behavior in

animals, but virtually non-metabolized property of PFOA or PFOS ignores chances that

63

may result in observed different pharmacokinetics in two studies. Secondly, exposure
levels were far different in the two studies, which could modify removal kinetics. At the
highest treatment in a current study, blood PF OS concentrations were realized at most
tens of ppb level that were approximately three order less than those measured at tens of
ppm level in the previous study. That high exposure scheme was attributed to the fact
that they designed their experiment to determine the acute toxicity of dietary PFOS in
bird species. A biphasic type of PFOS elimination was seen in a mammalian study with
PFOS-treated cynomolgus monkeys (Seacat et al., 2002). In their study, the elimination
rate in first 50 days after cessation appeared to be faster than that at the end of a l-year
depuration period. Similarly, high PFOS-treated chickens eliminated blood-bome PFOS
almost twice as faster as did low PFOS-dosed chickens in this study. These observations
may indicate that the biological mechanism of PF OS elimination is acting differently at

lower— and higher— body burdens.

Table 9. Elimination half-life (days) of PFOA and PFOS in blood or serum from
other studies

 

 

 

 

PFOA Sex Dose Form Half-life Sources

Chicken Male Implant 5.2 This study

Rat Male i.v. 5.6 Ohmori et al. (2003)
Dog Male iv 20 and 23 Hanhijarvi et al. (1988)
Monkey Male i.v. 2l Noker (2003)

Human Both Occupational 1600 Burris et a1. (2002)
PFOS Sex Dose Form Half-life Source

Chicken Male Implant 125 This study

Mallard Both Dietary 6.9 Newsted et al. (2005)
Rat Male i.v.('4C-labelled) >90 Jonhson et al. (1979)
Monkey Male Oral lOO~200 Seacat et al. (2002)
Human Both Occupational 1428 BM (2000)

 

64

 

Chapter 3

Perfluoroalkyl Acids in Marine Organisms from Lake Shihwa, Korea

Abstracts

Concentrations of eight perfluoroalkyl acids (PFAs) were determined in three marine
fishes (mullet, shad, and rockfish) and marine invertebrates (blue crab, oyster, and
mussel) from Lake Shihwa, Korea, where great concentrations of PFAs in waters of Lake
Shihwa and adjacent industrial effluents were reported. Perfluorooctanesulfonate (PFOS)
was the dominant compound of PFAs analyzed in these marine organisms and most
PFOS concentrations were greater than the sum of all other PFAs monitored (PF HS +
ZPFCAS + N-EtFOSAA). The mean concentrations of PFOS were 8.1x 101 and 3.6><lOl
ng/g w in liver and blood of fish, respectively. Perfluorocarboxylic acids (PFCAs)
were also found in fish at ten-fold lower concentrations. Longer chain PFCAs, such as
perfluorododecanoic acid (PFDoA, C12) and perfluoroundecanoic acid (PFUnA, C11)
occurred in fish at greater concentrations than did perfluorononanoic acid (PFNA, C9)
and perfluorooctanoic acid (PFOA, C8). A spatial comparison of blue crab data showed
that PFOS concentrations in soft tissues decreased as a function of distance from the
shore and the inputs from the industrialized areas where wastewaters are discharged into
Lake Shihwa. Only PFOS was a detectable PFA in mussel and oyster with a mean of
0.5d20.2 ng/g and 1.1i0.3 ng/g ww, respectively. A significant positive correlation was
observed between concentrations of PFUnA and PFDA in both liver and blood of fish,
implying the same sources of PF CAs in this area. The calculated hazard quotients (HQs)

using the estimated whole body PFOS residue were less than 1.0 for all fishes. For

65

example, an HQ calculated from the greatest PF OS in fish was estimated to be only 6X10'
4. Current concentrations of PFOS in fish in Lake Shihwa are not expected to cause acute

lethality of the fish.

Introduction
Perfluoroalkyl acids (PFAs) have been commonly used in industrial applications since
the mid 1940’s, as surface protectors for carpets and leather, active-component in fire-
fighting foams and as processing aids in the production of fluorinated polymers (Kissa,
2001; Giesy and Kannan, 2002). Recent global PFAs monitoring of biota have found
that the perfluorooctanesulfonate (PFOS) and other related PFAs are ubiquitously found
in various tissues of animals ranging from invertebrates, fishes, birds, and mammals
globally irrespective of locations and some of them are biomagnified in the upper tropic
level organisms (Vijver et al., 2003; Sinclair et al., 2006). To date, PFOS has been found
to be the predominant compound among PFAs analyzed with current analytical
techniques, while another toxicological counterpart of interest, perfluorooctanoate
(PFOA) is not persistent in tissues of wildlife except human serum (Houde et al., 2006;
Kannan et al., 2004). While there have been some assessments of the ecological risks of
these newly emerged organic contaminants based on accumulated PFAs exposure data
around the globe (OECD, 2002; US. EPA, 2003), few data are available for assessing
current potential risks of PF As in wildlife from Korea (Kannan et al., 2002).

Lake Shihwa, located on the west coast of Korea, is an artificial saltwater lake
that has been receiving industrial wastewater discharges from the Shihwa and Banweol

Industrial Complexes (SBICs, approximate total industrial area = 31 kmz) since its

66

construction in 1994 (Figure 9). Although this man-made lake’s function was originally
planned to supply waters to SBICs and nearby agricultural areas, gradual deterioration of
lake water quality prompted scientific studies evaluating the environment quality of Lake
Shihwa (Jung et al., 1997). Those research efforts indicated moderate-to-relatively great
concentrations of trace metals and persistent organic pollutants such as PCBs, PAHs,
organochlorine insecticides, and alkylphenols in waters and sediments from this region
(Khim et al., 1999; Li et al., 2004). Recently, concentrations of PFAs in waters of the
lake and tributary streams were reported with greatly elevated water-borne PFOS,
suggesting that local sources discharge PFAs or their precursors into the lake
(Rostkowski et al., 2006).

Lake Shihwa (surface area = 56.5 kmz, drainage basin = 476.5 kmz) provides
valuable habitats for diverse aquatic animals and birds, particularly migratory species.
Because significant quantities of PFAs were observed in Lake Shihwa, it was deemed
prudent to determine concentrations of PFAs in aquatic animals of Lake Shihwa and
assess the potential risks that these compounds might pose to those organisms and the
organisms that might eat them. In the present study, concentrations of PFOS and other
PFAs were measured in various marine organisms (fish, crab, mussel, and oyster) from
Lake Shihwa. Since PFAs have a propensity to bind to proteins and be retained in the
blood and liver, fish blood and samples of liver and or hepatopancreas were collected at
various locations in Lake Shihwa for instrumental analysis. Second, the current status of
PFOS contamination in this region was assessed by comparing results from this study to
concentrations of PFOS that have been reported to occur in fish from other parts of the

world. Third, the sources of PFAs in this region are discussed based on the current

67

understanding of the environmental fate of PFAs. Finally, a screening level hazard

assessment was conducted using PFOS concentrations in the organisms.

Materials and Methods

Sample Collection

Fishes and blue crabs were collected in May 2006 at two locations near to the SBICs and
at two locations near the inland discharges from Lake Shihwa (Figure 9). Stationary fish
nets were installed at sampling locations for three days before collection. Immediately
after net fishing, all samples were kept in iceboxes and transported to the laboratory
where samples of blood and liver were taken. Unfortunately, large amounts of entrapped
jelly fish caused mortality of several fish species and resulted in only one or two fish
species being available at each location. Composite liver and blood samples (two fish per
composite, number of composite = 3) of mullet (Chelon haematochelia, mean fish weight
¥ 3.5X10'i4.5 g wet weight, ww), rockfish (Sebastes schlegeli, mean fish weight =
8.5><10':1:1.l><10l g ww), and shad (Konosirus punctatus, mean fish weight = 3.8X10'i6.0
g ww) were prepared. Fish blood was collected by use of a heparinated syringe and
transferred in a lSmL polypropylene (PP) tube. Mussel and oyster were collected at one
location from a barge ship located in the middle of the lake (Figure 9). Soft tissues of
blue crab (Portunus trituberculatus, three individuals at each location) and marine
invertebrates (mussel for n=6 and oyster for n=4) were used for PFAs analysis. Tissues
were then mechanically homogenized with a vortex mixer and kept at negative 20°C until
the extraction. To avoid cross contamination, the homogenizer probe and parts were

thoroughly rinsed with Milli-Q water and methanol (MeOH) between homogenizing.

68

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69

Chemicals and Standards

A mixture of PFAs was used as an external standard and contained two perfluoroalkyl
sulfonates that included PFOS and perfluorohexanesulfonate; PFHS) as well as five
perfluoroalkyl carboxylic acids including perfluorododecanoic acid (PFDoA);
perfluoroundecanoic acid (PFUnA); perfluorodecanoic acid (PFDA); perfluorononanoic
acid (PFNA); PF OA, and N-ethyl perfluorooctane sulfonamide acetate (N-EtFOSAA) at
10 ng/mL for each standard. The standards mixtures were supplied by the National
Institute of Advanced Industrial Science and Technology (AIST), Tsukuba, Japan.
['3C]PFOS and ['3C]PFOA were used as internal standards. Pesticide-grade MeOH,
ammonium acetate, and ammonium solution (25%) were purchased from Wako Pure

Chemicals, Japan. Milli-Q® (Bedford, MA) water was used in the whole experiment.

Sample Preparation

Biological samples were extracted with an ion-pairing liquid extraction method with
some modifications (Kannan et al., 2004). Briefly, an aliquant of blood (0.5 mL) or
tissue homogenate (0.5 g wet wt) was diluted 5-fold with Milli-Q water. One milliliter of
diluted blood or tissue-water mixture was then transferred into a 15mL PP tube.
Extraction solutions and solvent were pipetted into the mixture in the following sequence;
lmL of 0.5M ion-pairing agent (tetrabutyl ammonium hydrogensulfate) adjusted to pH
10, two milliliter of 0.25M extraction buffer (sodium carbonate + sodium bicarbonate),
and SmL of methyl tert-butyl ether (MTBE). Samples were rigorously shaken for 20 min
followed by centrifugation for 15 min (2><103 rpm). The isolated organic phase was

transferred to a clean 15 mL PP centrifuge tube. This extraction was repeated twice and

70

all combined MTBE supematants were evaporated to near dryness under a gentle stream
of nitrogen. The sample was then re-dissolved in 1 ml MeOH. Further clean-up of
resulting extracts were conducted using solid-phase extraction (SPE) with some
modifications (So et al., 2006). Briefly, a half milliliter aliquant of unfiltered extracts
was diluted with lOOmL Milli-Q water and consequential water-extract mixture was
passed through an Oasis WAX® cartridge (0.2g, 6 cm3) at an elution rate of l drop/sec.
Afier sample loading was completed, cartridge was washed in the sequence of buffer
(25mM acetic acid l70mL + 25mM ammonium acetate 30mL) adjusted to pH 4 and
MeOH. The fraction containing PFAs of interest was retrieved by eluting 0.1% NH4OH
dissolved in MeOH. Teflon or glassware was avoided in whole extraction procedure to

remove possible contamination of samples and sorption of analytes.

PFAs Analysis by LC/MS/MS and QA/QC

Separation and quantification of PFAs in biological samples were performed using an
Agilent HP 1100 liquid chromatography (HPLC) (Agilent, Palo Alto, CA) interfaced
with a Micromass Quattro H mass spectrometer (MS/MS) (Waters Corp., Milford,
Massachusetts) operated in electrospray negative ionization mode. Ten microliters
aliquot of extract was injected onto a Keystone Betasil C13 column (2.1 mm id. X 50 mm
length, 5 pm) with 2 mM ammonium acetate and MeOH as mobile phase, starting at 10%
MeOH. Detailed instrumental parameters can be found in Taniyasu et al., 2003.
Concentrations of PFAs in extracts were quantified using external calibration curves
(1X10'2, SXIO'Z, 2X10", IXIOO, 1><10l ng/mL). Acquired data were deemed to be

acceptable if QC standard measured afier every 10 injections fell within 30% of its

71

theoretical value, otherwise samples were run again with a newly constructed calibration
curve. The limit of quantification (LOQ) in this study was determined to be 0.25 ng/g
wet wt for all analytes in all sample types.

To ensure the quality of analytical procedures, procedural blank, procedural
recovery were performed with each set of extraction batch, and matrix-spiked recovery
were tested for each type of biota. All target PF As in procedural blanks were below LOQ.
Mean recoveries (n=2) from each biota homogenate sample (0.5 g or 0.5 mL) spiked with
10 ng PFAs were 90.1% for PFOS, 87.5% for PFHS, 81.1% for PFDoA, 97.0% for
PFUnA, 105.5% for PFDA, 100.9% for PFNA, 80.4% for PFOA, 79.5% for ['3C]PF0A
and 78.1% for N-EtFOSAA. Mean matrix-spike recoveries for two internal standards
[13C]PFOS and ['3C]PFOA were 90.6% and 78.1%, respectively. Concentrations
reported in this paper were not adjusted for matrix recoveries. For calculation of mean
concentrations, values determined to be less than the LOQ were assigned as one half of

LOQ.

Data and Statistical Analysis

The normality of the concentration data was analyzed by means of a Kolmogorov-
Smirnov test. Differences in PP OS concentrations in tissues of animals among locations
were tested by one-way ANOVA (Type I error level of 5%). The Tukey test was used as
post-hoe criterion. Because equal variance was not obtained for PFOS concentrations
from Lake Shihwa, Japan, and other locations, to which concentrations determined in this
study, were compared by use of the non-parametric Kruskal-Wallis test without data

transformation. Associations between PFA concentrations in fish were investigated using

72

a Pearson correlation analysis. All statistical analyses were performed with the

SYSTAT® 11 Package (SYSTAT Software Inc., Richmond, CA).

Results

PFAs Concentrations in Marine Fishes

Concentrations of two sulfonate PFAs (PFOS and PFHS), and homologous carboxylic
PFAs with carbon-chain lengths between 8 and l2, and N-EtFOSAA were determined in
blood and liver of three marine fish species (rockfish, shad, and mullet) (Table 10). As
has been reported in previous studies of PFA in biota, PFOS was the predominant PFA in
fish caught in Lake Shihwa. Concentrations of PFOS in all individual fish were greater
than the sum of all other PFAs monitored (PFHS + ZPFCAs + N-EtFOSAA). In general,
fish liver contained greater concentrations of PFOS than did blood (12 out of 18
individuals). A sample of liver from a mullet captured at station 4 had 2.5><102 ng
PFOS/g ww, which was the greatest concentration of PFAs determined in this study;
meanwhile blood samples from mullet at station 3 contained a concentration as great as
9.3><10l ng PFOS/mL. Analyzed by species, mullet (l.9><102i3.2><10l ng/g, ww)
accumulated 6.8 to 8.2-fold greater concentrations of PFOS in their livers than did
rockfish or shad (Figure 10). In contrast, the fold-difference in concentration of PFOS in
blood ranged from 1.9- (mullet to rockfish) to 3.0-fold (mullet to shad). PFHS was
detectable in most samples and concentrations in liver (<LOQ-2.0 ng/g ww) and blood
(<LOQ-l.6 ng/mL) were similar. Detectable concentrations of N-EtFOSAA were

measured in livers of mullet (14.1i4.3 ng/g ww), however in rock fish and shad, N-

73

EtFOSAA was in the range of <LOQ to 1.8 ng/g w in liver and <LOQ to 4.2 ng/mL in

blood.

74

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In general, longer chain carboxylated PFAs, such as PFDoA (C12) and PFUnA
(Cll) occurred in fish at greater concentrations than did PFNA (C9) and PFOA (C8)
(Table 10). Liver of mullet contained relatively great concentrations of PFUnA
(2.5><10'i4.5 ng/g ww) which was the greatest concentration of the PFCAs, followed by
PFDoA > PFDA > PFNA > PFOA. Concentrations of PF CAs in rockfish and shad were
only detectable in the liver (<1 .5 ng/ g w). The mean hepatic PFOA concentration was
1.2:t0.7 ng/g ww in all fishes. As observed for PFOS, lesser concentrations of PFCAs
were measured in blood than liver of mullet. The greatest concentration of PFCA in fish
blood was observed for mullet (1.1X10'i2.7 ng PF UnA/mL) with lesser concentrations in
blood observed in decreasing of PFDA > PFDoA > PFNA > PFOA. In contrast,
concentrations of PFCAs in blood of other fish species were less than those in mullett,
with concentrations ranging from 1.6 to 3.6 ng/mL in rockfish and from 0.6 to 1.4 ng/mL

in shad.

PFAs Concentration in Marine Invertebrates

Mean concentrations of PFOS, PFHS, PFDoA, PFUnA, PFDA, PFNA, PFOA and N-
EtFOSAA in marine invertebrates are summarized in Table 11. While blue crabs were
available at four locations (Station l~4), mussels and oysters were only available in the
middle of the lake (Station 5). Similar to PFAs accumulation patterns in fish, PFOS was
the predominant PFA analyzed in these marine organisms. PFOS concentrations in the
soft-tissue of crabs from Station 3 (9.6i1.0 ng/g ww) and Station 4 (8.3i0.l ng/g ww)
contained significantly greater PFOS concentrations than those from Stations 1 and 2

(p<0.05). PFOS concentrations in mussel and oyster collected from Station 5 were

77

0.6i0.2 ng/g ww and 1.1:l:0.3 ng/ g ww, respectively and were less than those observed in
blue crabs sampled in this study. In blue crab samples, concentrations of PFDoA ranged
from 1.4 to 6.2 ng/g ww while PFUnA concentrations ranged from 1.4 to 1.9 ng/g ww.
The concentrations of PF DoA and PFUnA were greater than the other quantified PFCAs,
which contained concentrations ranging from 0.3 to 0.8 ng/g ww. PFOA was detected in
all crab tissues but at quantifiable concentrations that ranged from 0.25 to 0.82 ng/g ww.
With the exception of PFOA in oysters, all concentrations of the targeted PFCAs in
mussel and oyster were less than their respective LOQ values. No quantifiable
concentrations of either PF HS or N-EtFOSAA were observed in crabs, mussels or oysters

(<LOQ).

78

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Profiles of Relati ve Concentrations of PFAs in Marine Organisms

Profiles of relative concentrations of PFA in the three species of fish included in this
study were similar to those reported in other studies (Houde et al., 2006), with PFOS
accounting for the greatest portion of the total mass of PFAs (Figure 11)., The
proportions of PFAs contributed by PFOS ranged from 72.0~85.2% in liver and
64.1~77.6% in blood. The next most prevalent PFAs were the longer chain PFUnA,
which accounted for 5.4»~12.9% and 4.2~9.3% of the total PFA concentrations in blood
and liver, respectively.

While PFOA is ubiquitous in water samples and is also one of the perfluoroalkyl
chemicals found in human blood (Yamashita et al., 2004; Kannan et al., 2004), its
contribution to the total blood concentration of PFAs in fish was less than 2.5%. The
ratios of PFOS to total PFCAs (EPFCA) concentrations in liver ranged from 3.2 in mullet
to 6.5 in shad while for blood the ratios ranged from 2.1 (mullet) to 4.1 (shad). In blue
crabs, the ratio of PFOS to total PFCA in ranged from 0.9 in samples collected from

Station 2 to 2.0 in samples collected from Station 3.

80

 

 

 

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Discussion

Global Comparison of PF OS Concentrations

Extensive monitoring efforts have surveyed PFAs in biological matrices beginning in
2001 and delineated the extent of PFAs concentrations in wildlife, especially in marine
mammals in north-America and northern Europe (Giesy and Kannan, 2001; Houde et al.,
2006). In Korea, levels of PFAs in waters were first studied in 2005 in the streams of the
Shihwa-Banweol industrial complex and Lake Shihwa (Rostkowski et al., 2006). To our
knowledge, this is the first report regarding the concentrations of PFOS and other
perfluoroalkyl compounds in marine organisms from Korea. Current concentrations of
PF OS in the biota of Lake Shihwa were compared to concentrations of PF OS in fish from
other locations around the world (Figure 12). In marine fishes collected around Japan,
the median PFOS concentration was 6.4X10' ng/g w in livers (Taniyasu et al., 2003).
Global concentrations of PFOS in fish liver (excluding Japan) were also compiled and
compared to livers of fish from Lake Shihwa. The global range in PFOS concentrations
was 5.3 to 4.1X10' ng/g ww with a median of 1.4X10' ng/g ww. Fishes from Lake
Shihwa had a median concentration of 3.1X10' ng PFOS/g ww liver. This value was
significantly greater than the median concentration of PFOS in fish liver globally, but
was not statistically different from the median in Japan (Mann-Whitney U test, p=0.05).
In the case of fish blood, there was no statistical difference between the median for Lake
Shihwa (2.7><10I ng PFOS/mL) and that for Japan (3.4X10' ng PFOS/mL). Fishes
collected from the Mississippi River in proximate to the 3M Cottage Grove contained far
greater PFOS in (blood than Lake Shihwa fishes did (MPCA, 2006). For example, a

concentration of blood sample of smallmouth bass ranged from l.7><lO3 to 4.2><103 ng

82

 

PFOS/mL. However, since we were not able to locate additional data on PFOS
concentrations in blood of fishes from other global locations, meaningful comparisons
could not be made. The relatively great PFOS concentrations observed in fish from Lake
Shihwa when compare to the global data'may be due to the large amounts of industrial

and domestic wastes that Lake Shihwa receives on an annual basis.

83

 

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84

Source Estimation of PFAs in Lake Shihwa

The spatial distribution of PFOS concentrations in marine organisms demonstrates that
biota samples collected from sites close to the outlets of inland creeks (Stations 3 and 4)
were more contaminated than those sampled at sites away from the release sources
(Stations 1 and 2) (Table 10 & 11). The greatest mean concentration of PFOS in livers of
fish from Lake Shihwa was measured in mullet sampled at Station 4 (2.1 X102 :I: 3.8><1OI
ng/g ww liver), while concentrations were approximately lO-fold less in livers of fish
collected at Stations 3 and 4. Due to the nature of anionic organic compounds to
generally be found in water column rather than to associate with particulates and
sediments, elevated concentrations of PFOS in fish and blue crab tissue were primarily
attributed to exposure to PFOS in the water column. Concentrations of PFOS in the
waters at Stations 3 and 4 were reported to be 1.3X10' and 1.8X10' ng PFOS/L,
respectively (Rostkowski et al., 2006). Although concentrations of PFOS in water
around Station 1 and 2 are not available, previously reported concentrations of PFOS at
locations fiirther from the known effluents would indicate that they should be less at these
two stations. In addition, Station 3 and 4 are close to wastewater releases from inland
channels containing greater concentrations of PFOS (2.O><10'-6.5><102 ng PFOS/L in
streams receiving effluents from SBICs). Direct comparison among locations is also
complicated by the fact that different species of fish may have different rates of PFOS
accumulation. For example, PF OS concentrations in shad did not differ between Stations
1 and 3. However, a comparison of blue crab data collected at all four locations showed

that PFOS concentrations in soft tissues decreased as a function of distance from the

85

shore and the inputs from the urbanized and industrialized areas where wastewaters are
discharged into Lake Shihwa (Table 11).

Even though PF 0A was the second most abundant PFA in seawater samples (4.5-
1.1X10' ng/L) collected from Lake Shihwa, PFOA concentrations in fish were among the
least of the targeted PFAs measured in this study. This observation was similar to that
reported for other studies (Martin et al., 2004). In general PFOA has a lesser
bioaccumulation potential than other longer chain PFCAs and sulfonated PFAs (Martin et
al., 2003a, b). The occurrence of PFOSA in fish was reported elsewhere (Houde et al.,
2006, Sinclair et al., 2006), but was not quantified in our samples. However, detection of
N-EtFOSAA in marine biota suggests that precursor compounds for PFOS exist in this
region and therefore, to a lesser degree, contribute to the overall accumulation of PFOS
by marine fishes.

A significant positive correlation was observed between concentrations of PFUnA
and PFDA in both liver and blood of fish (Figure 13). Fluorotelomer alcohols (FTOH)
are known precursors of PFCAs via several mechanisms such as atmospheric degradation,
hydrolysis, and biodegradation (Ellis et al., 2004; Gauthier and Mabury, 2005; Wang et
al., 2005). For example, 8:2 FTOH can be degraded to PFNA and PFOA, and through
the same processes, 10:2 FTOH could produce PFUnA and PFDA. Thus, significant
relationships between PFUnA and PFDA in fish samples are indicative of the existence
of the same source and in part, contamination of the semi-volatile 10:2 FTOH may be a
source of PFUnA and PFDA in the Lake Shihwa area. The observation of greater

concentrations of PFUnA than PFDA in fishes of Lake Shihwa are consistent with the

86

bioconcentration study results that PFUnA is more bioaccumulative than PFDA (Martin

et al., 2003a).

BCFs for PFAs in Fish

Bioconcentration factors (BCF) were estimated for fish by dividing the
concentrations of PFAs in fish by those in seawater, which had been previously
determined for Lake Shihwa. Though bioaccumulation factor (BAF) is a more accurate
estimate for describing the bioavailability from a field exposure, herein we used BCF
terminology for convenience. However, because of a lack of information on PFA
concentrations in seawater in the outer lake (Station I and 2), and the fact that the more
bioaccumulative, longer chain PFCAs (PFDoA and PFUnA) were not measured in waters
of Lake Shihwa, it was not possible to estimate BCFs for these PFAs. BCF values were
determined only for PFOS, PFHS, PFDA, PFNA, and PFOA in mullet collected at
location 3 and 4. Values of the BCF for PFOS were estimated to range fiom 2.2><103 to
5.9><103 and l.l><104 to 1.4x 104 in mullet blood and liver, respectively. These ranges are
similar to those reported for fishes in Tokyo Bay (1.4X103-2.1X104) (Taniyasu et al.,
2003), but less than BCF values reported for fish from an area where a spill had occurred
in Etobicoke Creek in Toronto (6.3X103-1.3><105) (Moody et al., 2002). BCF values for
PFOS determined under laboratory conditions were estimated to be 4.3><103:!:5.7><102 in
blood and 5.4><103.Jc8.6><102 in liver of rainbow trout “(Oncorhynchus mykiss) at the
steady-state condition (Martin et al., 2003a). BCF for the six carbon PFHS was less than
PFOS, with values ranging from 1.9><102 to 8.3><102 in blood samples and 6.4><1O2 to

1.5><103 in liver samples.

87

 

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The site-specific BCF values of PFCAs estimated for Lake Shihwa waters were
greater than those estimated under laboratory conditions. For example, the BCF value for
PFOA which is a ubiquitous PFA in water column was in the range of 5.8><10l to 3.8><102
in mullet samples, while laboratory-derived BCFs were only 8 to 2.7><lOl in rainbow
trout. Field-determined BCF for PFDA (6.9X103 in blood and l.5><104 in liver) in this
study was also greater than those estimated in rainbow trout under laboratory condition
(2.7><103 and l.l><103 for blood and liver, respectively). Each fish species may have
different toxicokinetics for PFCA accumulation, which results different BCF. However,
this observation may suggest the existence of unidentified precursors of PFCAs in the
region of Lake Shihwa, which makes significant contribution to the overall

bioaccumulation of middle-length PF CAs (C8-C12) in marine organisms.

Hazard Assessment based on Concentrations of PF OS in Tissues

An earlier hazard assessment based on water-bome PFOS concentrations revealed that
neither waters inland streams nor lake Lake Shihwa waters exceed water-quality
(guideline) values that are protective of aquatic species (Rostkowski et al., 2006).
Concentrations of PFOS in tissues determined in the present study were used to evaluate
the ecological risk of PFOS exposure to fish in Lake Shihwa. The tissue concentration
that would not be expected to cause acute effects in fish was determined to be 8.7><10l
mg PFOS/kg ww, which is less than the 95% confidence limit of the LDOI based on a
study of the blue gill sunfish (Beach et al., 2005). Prior to hazard quotient (HQ)
calculation, the hepatic PFOS concentrations were converted into whole carcass PFOS

concentrations. To conduct the risk assessment based on whole body concentrations a

89

conversion ratio was developed by dividing the concentration in liver to that in whole
body. The value of 4.9 was used to predict the whole body concentration and then a
threshold water concentration was estimated by using the BCF value for PFOS that was
determined from a water-only exposure of rainbow trout (Martin et al., 2003a). By using
this conversion, the whole body concentrations of PFOS in fish were approximated be in
the range of 3.9 to 5.1x10l ng/g, w. The calculated HQs were less than 1.0 for all fishes.
For example, an HQ calculated from the greatest PFOS in fish was estimated to be only
6X104. These results, estimated fi'om the measured concentrations in fish tissue were
consistent with those predicted previously from measured water concentrations by use of
a bioconcentration factor (So et al., 2004) that least current concentrations of PFOS in
fish in Lake Shihwa are not expected to cause acute lethality of the fish. However,
currently there is little information on the potential for chronic effects of PFOS on fish in

Lake Shihwa or to assess the potential for effects of the mixture of PFAs.

9O

Chapter 4

Perfluoroalkyl Acids in the Egg Yolk of Birds from Lake Shihwa, Korea

Abstract

Concentrations of perfluoroalkyl acids (PFAs) were measured in egg yolks of three
species of birds collected in and around Lake Shihwa, Korea, which receives wastewaters
from an adjacent industrial complex. Mean concentrations of perfluorooctanesulfonate
(PFOS) ranged from l.9><102 to 3.l><102 ng/g ww, with size-dependent accumulation
among species. Measured concentrations of PFOS were similar to those reported for bird
eggs from other urban areas. Concentrations of long-chain perfluorocarboxylic acids
(PFCAs) were found in egg yolks. Mean concentrations of perfluoroundecanoic acid
(PFUnA) in egg yolks of little egret (Egretta garzetta), little ringed plover (Charadrius
dubius), and parrot bill (Paradoxornis webbiana) were 9.5X10', 1.5X102, and 2.O><102
ng/g ww, respectively. Perfluorooctanoic acid (PFOA) was detected in a few samples,
but concentrations were lOO-fold less than those of PFOS. Relative concentrations of
PFAs in all three species were similar, with the predominance of PFOS (45-50%)
followed by PFUnA (25-30%). There was a statistically significant correlation between
PFUnA and perfluorodecanoic acid (PFDA) in egg yolks (r2=0.54). This suggests that
fluorotelomer alcohols are important contributors to the occurrence of long chain PFCAs
in bird eggs. Using measured egg concentrations as well as concentrations in the diet of
the birds, ecological risk of PFOS and PFAs to birds in Lake Shihwa was evaluated using
two different approaches. Estimated hazard quotients were similar between the two

approaches. The concentration of PFOS associated with 90th centile in bird eggs was

91

lOO-fold less than the lowest observable adverse effect level (LOAEL) determined for
birds and those concentrations were 4-fold less than the suggested toxicity reference
values (TRV). Current concentrations of PFOS are less than what would be expected to

have an adverse effect on birds in the Lake Shihwa region.

Introduction

Widespread occurrence of perfluoroalkyl acids (PFAs) in wildlife has spurred monitoring
efforts and regulatory concerns regarding these emerging contaminants (Giesy and
Kannan, 2001&2002; OECD, 2002; Environment Canada, 2006). The physico-chemical
properties of PFAs make them very useful for application in various commercial products
such as surface protectors for carpets and leather, active-components in fire-fighting
foams, and processing aids in the production of fluorinated polymers (Kissa, 2001). Food
web studies and examination of concentrations in biota suggest that
perfluorooctanesulfonate (PFOS) and other related PFAs are bioaccumulative to some
extent and are able to biomagnify to top-predators such as marine mammals and fish-
eating birds (Kannan et al., 2001; Van de Vijver et al.; Martin et al., 2004; Houde et al.,
2006). PFAs and in particular PFOS and perfluorooctanoate (PFOA), which were once
considered to be biologically inert are relatively bioactive at the cellular level, causing
diverse effects including blockage of cell-cell communication (Upham et al., 1998) and
initiation of hepatic peroxisome proliferation (Intrasuksri et al., 1998). These two PFAs
have also been shown to cause developmental toxicities in experimental animals
including rodents, birds, and amphibians (Austin et al., 2003; Lau et al., 2004; Molina et

al., 2006; Palmer and Krueger, 2000).

92

Birds from urbanized areas contain greater concentrations of PFAs in their tissues
or egg than those from more remote areas (Kannan et al., 2001; Verreault et al., 2005;
Sinclair et al., 2006). Concentrations of PFAs in osprey eggs were also correlated with a
concentration gradient of other persistent contaminants in the environment suggesting
local sources of exposures (Toschik et al., 2005). Furthermore, guillemot eggs collected
from 1968 to 2003 showed increase in PFOS concentrations by 30-fold from 1968
(2.5X10l ng/g ww) to 2003 (6.l><102 ng/g ww), which corresponds to greater use of
fluorochemicals (Holmstrom et al., 2005).

Lake Shihwa, located on the west coast of Korea is an artificial saltwater lake that
has received industrial wastewater discharges from bordering Shihwa and Banweol
Industrial complexes (SBICs, approximate total industrial area = 31 kmz) since its
construction in 1994 (Figure 14). Investigation of trace metal and persistent organic
pollutants, including PCBs, PAHs, organochlorines, and alkylphenols in water and
sediment from Lake Shihwa and its neighboring industrial complexes, have suggested a
moderate-to-high degree of contamination (Khim et al., 1999; Li et al., 2004).
Concentrations of PF As of waters of Lake Shihwa and creeks running through the SBICs
and lake organisms such as fish and marine invertebrates have recently been reported
(Rostkowski et al., 2006; Yoo et al., submitted). These earlier studies found relatively
great concentrations of PFOS in the water column and certain fishes and invertebrates.
Concentrations of PFAs in livers of birds from other areas of Korea have been reported
on limited sample sizes (Kannan et al., 2002). However, no attempt has been made to
investigate the concentrations and effects of PFAs in the upper-tropic organisms, such as

fish-eating birds in this region.

93

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The aim of this study was to measure concentrations of PFOS and other PFAs
eggs of birds collected in the vicinity of Lake Shihwa. Egg yolks were selected for this
study because the concentrations of PFAs in eggs are often used in exposure and risk
assessments. Assessment of the potential risks of PFOS and a mixture of PFAs to birds
in the Lake Shihwa area were made based on the currently available benchmark doses for
toxicity in birds (Newsted et al., 2005). Concentrations of PFAs in water and fish-diet of
birds were also incorporated to provide multiple lines of evidence for assessing risk. In
biota, PFAs were present as a mixture of PFAs rather than a single congener. Thus, he
gap junction intercellular communication (GJIC) cell-bioassay was used to obtain a
relative toxic potency of individual PF As to calculate PF OS-equivalent concentrations in

egg samples for the risk assessment of PFA mixture (Villeneuve et al., 2000).

Materials and Methods

Egg Collection and Sample Extraction

Eggs of three species of birds were collected from the area in and around Lake Shihwa
during the breeding season of May 2006 (Figure 14). In this area, one colony of little
egret (Egretta garzetta) was found in a section of the city of Ansan, while eggs of little
ringed plover (Charadrius dubius) were sampled at various locations including the
islands. Nests of parrot bill (Paradoxornis webbiana) occurred in wetlands located
upstream of Lake Shihwa. One or two eggs per nest were collected and a total of ten,
seven, and four nests were surveyed for little egret, little ringed plover, and parrot bill,
respectively. Mean shell length (mm) and egg weight (g) of each species are summarized

(Table 12). To avoid potential contamination during sampling, eggs were stored

95

individually in a polypropylene container and kept at -20 °C until instrumental analysis.
Egg yolks were extracted with an ion-pairing liquid extraction method described
elsewhere (Kannan et al., 2004). Briefly, an aliquant of homogenized egg yolk (1 g ww)
was diluted 4 fold (w/v) with Milli-Q ‘water before extraction. Five nano-grams of
['3C]PFOS and [‘3C]PFOA were spiked as internal standards. Tetrabutyl ammonium
hydrogensulfate (TBA) and methyl tert-butyl ether (MTBE) were used as an ion-pairing
agent and extraction solvent, respectively. The resulting extracts were concentrated
under a gentle stream of nitrogen and reconstituted in 1.0 mL methanol. Teflon or
glassware was avoided in whole extraction procedure to remove possible contamination

of samples and sorption of analytes.

96

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Instrumental Analysis and QA/QC

Target analytes of perfluoroalkyl acids in extracts were quantified using an Agilent 1100
series high-performance liquid chromatograph (HPLC) coupled with an Applied
Biosystems API 2000 tandem mass spectrometer (MS/MS). The extract was injected
separately for sulfonate and carboxylic acids. For sulfonate acids, 10 % methanol at a
flow rate of 300 ul/min was increased to 100 % methanol at 10 min and was held for 5
min and then was back to 10 % methanol. For carboxyl acids, 100 % methanol was used
at 7 min and was held for 3 min. The MS/MS was operated in electrospray negative ion
mode. Analyte ions were monitored in multiple-reaction monitoring mode. Parent and
daughter ion transitions used for identification and quantification were 399>80
(perfluorohexane sulfonate, PFHS), 499>99 (perfluorooctane sulfonate, PFOS), 498>78
(perfluorooctane sulfonamide, PFOSA) for sulfonate acids and 413>369
(perfluorooctanoic acid, PFOA), 463>419 (perfluorononanoic acid, PFNA), 513>469
(perfluorodecanoic acid, PFDA), 563>519 (perfluoroundecanoic acid, PFUnDA),
6l3>569 (perfluorododecanoic acid, PFDoDA) for carboxylic acids. Two internal
standards of 503>99 (”C-PFOS) and 417>372 (”C-PFOA) were also monitored for
analytical recovery of sulfonate and carboxylic acids, respectively. Concentrations of
PFAs in extracts were quantified using an extracted calibration curve containing target
analytes spiked into a chicken egg matrix (0.1, 0.2, 0.5, 1.0, 5.0, and 20.0 mg/L). The
chicken egg was found not to contain measurable concentrations of target compounds.
The coefficient of determination (12) for each constructed curve was greater than 0.99.
Acquired data were deemed acceptable if QC standard measured after every 10 injections

fell within 30% of the long-term average, otherwise analysis was stopped and samples

98

 

were reanalyzed with a new calibration curve. The limit of quantification (LOQ) was
estimated at the lowest concentration point on the calibration curve within i30% of its
theoretical value. The LOQ in this study was determined to be 0.8 ng/g ww for all
analytes. Procedural blanks and matrix-spiked recovery tests were used to ensure the
quality of the analytical procedure. Concentrations of any target PF As in the procedural
blank (n=6) were below the LOQ. Mean recoveries (n=3) from egg yolk samples spiked
with 5.0 ng PFAs were 83.3% for PFOS, 90.0% for ['3C]PFOS, 68.0% for PFOSA 63.6%
for PFHS, 151.2% for PFDoA, 157.1 for PFUnA, 128.2% for PFDA, 128.6% for PFNA,
68.4% for PFOA, and 79.5% for ['3C]PFOA. For calculation of mean concentrations,

values below LOQ were arbitrarily assigned as half of LOQ.

Gap Junction Intercellular Communication (GJIC) Cell Bioassay

Rat liver epithelial cells (WB-F344 cells) were used to measure the inhibition potential of
cellular communication by individual PFAs. WB-F344 cells were obtained from Drs.
J .W. Brisham and M.S. Tsao of University of North Carolina. The cells were cultured as
described previously (Upham et al., 1998). GJIC was measured using the scrape loading
dye transfer technique (Hu et al., 2002). Briefly, confluent cells were trypsinized with 1><
trypsin-EDTA and cell solution was harvested. Two milliliters of the diluted cell solution
were seeded to 35 mm tissue culture plates, and cells were allowed 48 h for attachment
before the PFA dosing. Total twelve PFAs (PFBA, PFBS, PFHA, PFHS, 6:2 FTOH,
PFOS, PFOA, 8:2 FTOH, PFDA, 10:2 FTOH, 10:1 FTOH, and PFTetraDA) were tested.
At day 2, cells were exposed to each PFA dissolved in acetonitrile (0, 3.125, 6.25, 12.5,

25, 50, and 100 ug/mL) for 15 min. Following exposure, cells were rinsed with

99

phosphate buffered saline (PBS) and 1 ml of 0.05% lucifer yellow dye (Sigma, St. Louis,
MO) was added to each plate. A surgical steel blade was used to make three scrapes
through the monolayer of cells. After 5 min incubation at room temperature, the dye was
discarded, and the cells were rinsed with PBS, and then fixed with 0.5 ml of 4% formalin.
Dye migration was photographed at 200x using a Nikon epifluorescence phase contrast
microscope illuminated with an Osram HBO 200W lamp and equipped with a COHU
video camera. The area of dye migration from the scrape indicates the ability of cells to
communicate with each other through gap junction. The migrated area was calculated
using the Gel Expert program (Nucleotech, San Mateo, CA). Each PFA concentration

was tested in triplicate.

Data and Statistical Analysis

From a concentration-response relationship, EC50 of each PFA was calculated. Toxic
equivalent factor (TEF) was obtained by normalizing the EC50 of individual PFA
congener to that of the PFOS. PFOS equivalent concentrations (PEQ) in samples were
calculated by multiplying a corresponding TEF with its concentration in the sample.
Assuming additivity of toxic effects, a summed concentration of PEQs was used to
characterize the risks associated with a mixture of PF As in bird eggs.

To evaluate the differences in accumulation of PF OS and total PFCAs in egg yolk
among the bird species, one—way analysis of variance (ANOVA) was conducted with the
Bonferroni post-hoe criterion. Prior to analysis, values for concentrations of PFOS and
total PFCAs were log-transformed to attain the normality (one-sample Kolmogorov-

Smimov test) and equal variances (Levene's test). Probable associations among

100

perfluorinated carboxylic acids were tested using a Pearson correlation analysis. A probit
analysis was used to derive EC50 of inhibition of cellular communication (GJIC) using
Excel® program. Statistical significance was set at the level of p30.05, unless otherwise
noted. All statistical analyses other than EC50 determinations were performed with the

SYSTAT® 11 statistical package (SYSTAT Software Inc., Richmond, CA).

Results and Discussion

Concentrations in Birds

Mean concentrations of PFHS, PFOS, PFOSA, PFOA, PFNA, PFDA, PFUnA, and
PFDoA in egg yolks are summarized in Table 13. Significant amounts of PFOS were
detected in all the egg yolk samples. PFOA was measurable in a few samples at
concentrations two orders of magnitude less than that of PFOS. The long chain PFCA,
PFUnA was the second most prevalent perfluorinated compound in the egg samples.
Mean concentration of PFOS in egg yolk of little egret, little ringed plover, and parrot bill
was 1.9X102, 2.2X102, and 3.l><102 ng/g ww, respectively (Figure 15). The greatest
concentration of PFAs in this study was 1.2x 103 ng PFOS/g ww determined in an egret’s
egg yolk. In general, concentrations detected in bird eggs from the Lake Shihwa region
were similar to those seen in bird eggs and yolks from the Norwegian Arctic, Great Lakes,
and Delaware Bay, except for guillemot eggs from the Baltic Sea (5.6><102 - 8.7><102 ng
PFOS/g ww during 2000-2003 sampling campaigns) (Kannan et al., 2001; Verreault et al.,
2005; Holmstrom et al., 2005; Toschik et al., 2005). PFHS was also measured in some
egg samples, at concentrations as great as 9.5 ng/g ww. PFOSA was frequently found in

the diet of birds (fish samples), but were not quantifiable in any yolk samples (<LOQ).

101

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The concentrations of PFCAs in egg yolks of the birds analyzed were some of the
greatest ever found in wildlife species. In particular, the concentrations of long chain
PF CAs were elevated in all three species (Table 13). Among the PFCAs measured in this
study, PFUnA (C11) was the dominant compound followed by PFDoA (C12), PFDA
(C10), PFNA (C9), and PFOA (C8). Mean concentrations of PFUnA in little egret, little
ringed plover, and parrot bill were 9.5><10', 1.5X102, and 2.0><102 ng/g ww, respectively.
The predominance of PFUnA was also observed in osprey eggs from the Delaware River
and Bay. PFOA was detected in 60%, 100%, and 43% of little egret, plover, and red-
billed tit eggs, respectively. In plover, the concentration of PFOA was as great as
2.5><10l ng/g ww. In other studies, PFOA was not detectable in the eggs of guillemot and
glaucous gulls from northern Europe. Concentrations of PFCAs in eggs from our study
were lO-fold greater than those in glaucous gulls from the Norwegian Arctic (total
PFCAs was 4.2><10l ng/g ww with the dominance of PFUnA).

Mean concentrations of PFOS in birds were inversely proportional to length of
eggs. The smallest parrot bill egg contained the greatest concentration, while the largest
little egret egg contained the least concentration of PFOS in their yolks. However, this
difference was not statistically significant (Figure 15) Q)>0.05). The occurrence of great
mean concentrations of PFOS and total PFCAs in parrot bill is interesting because these
eggs were sampled at upper freshwater wetlands away from known sources such as
industrial complexes and populated areas, whereas little ringed plover eggs were
collected on islands and walkways in the vicinity of PFA contaminated Lake Shihwa.

Although the foraging ranges of investigated species were not studied, high

104

concentrations found in samples from wetland waters of Lake Shihwa need to be

investigated.

Contamination Sources to Birds and Relationship among PF CAs
Little egret and plover prey on aquatic organisms from Lake Shihwa and its watershed.
The exposure concentrations in these two species of birds are presumably an indication of
local PFA exposures. The existence of local sources pf PFAs in Lake Shihwa has been
suggested previously (Rostkowski et al., 2006). Concentrations of PFA in water were
elevated in drains and streams receiving effluents from the nearby Shihwa Industrial
Complex, and gradually decrease in Lake Shihwa with distance from the industrial
complex and were even less in the near-shore regions of Gyeonggi Bay. Concentrations
of PFOS and total PFCAs in fish and blue crabs also decreased as a function of distance
from wastewater discharges to the water exchange gate in Lake Shihwa (Yoo et al.,
submitted). Information on neither the amounts of PFAs used in the industrial complexes
nor their emissions from these industries bordering Lake Shihwa were not available, but
the results of previous studies have indicated that there are local sources in this area.
Profiles of relative concentrations of PFAs in the three species of birds examined
were similar to each other (Figure 16). The relative contribution of PFOS to the total
PFAs was less than what has been reported for other species, such as fish, reptiles, and
polar bears. Although PFOA concentration (1.7-1.1 ><10l ng/L) in waters of Lake Shihwa
was comparable to that of PFOS (7.3-1.8><10l ng/L), the relative concentration of PFOA
was less than 2% in egg yolks. PF UnA accounted for 25-30% of the total mass of PFAs.

In contrast, PFNA which is ofien a dominant PFCA in arctic marine mammals accounted

105

for only 10-20% of total PFCA concentration. Considering the differences in habitat,
location, body size, and diets among birds surveyed, the similarity of composition of
PFAs in eggs suggests that all three species have been exposed to a common source of
PFAs.

There was a statistically significant positive correlation between PFUnA and
PFDA (3:054), but not between PFNA and PFOA (r2=0.17) (Figure 17). There is
evidence that fluorotelomer alcohols (FTOHs) could generate PFCAs via atmospheric
oxidation, aqueous photolysis, and biodegradation (Ellis et al., 2004; Wang et al., 2005).
Studies have shown that 8:2 FTOH could degrade into PFNA and PFOA with even
number PFCA being a major product. Degradation of 10:2 FTOH has not been tested,
but presumably its degradation pathways may lead to PFUnA and PFDA. In most
samples of egg yolk concentrations of PFCA with an odd number carbon were greater
than the adjacent even number of carbon. For example, the slopes of correlation were 1.7
for PFUnA and PFDA, and 2.1 for PFNA and PFOA. This observation is explained by
differences in exposures and bioaccumulation potential among the PFCAs. Fish and
marine invertebrates from Lake Shihwa contained moderate to large concentrations of
PFCAs as well as PFOS (Yoo et al., submitted). In marine animals, concentrations of
PFCAs with an odd number of carbon atoms were greater than PFCA with an even
number of carbon atoms. For example, the concentrations of PFUnA in liver (mean, 9.2;
range, O.9-3.l><10l ng/g ww) were almost twice the hepatic concentrations of fish (mean,
4.6; range O.5-1.5><10l ng/g ww) collected from Lake Shihwa. Therefore, accumulation
pattern of PFCAs in birds from Lake Shihwa is influenced by PFCA concentrations in

prey items and BAF of individual PFCAs.

106

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Based on the current understanding of the degradation of FTOHs, a strong
relationship between PF UnA and PF DA in egg yolks suggests that 10:2 FTOH and/or its
unidentified precursors are likely contributors to the overall concentrations of long chain
PFCAs in the Lake Shihwa area. At present, however, we do not have atmospheric

concentrations of fluorinated contaminants in this region.

PF OS-Equivalent Concentration

Among twelve PFAs tested, eight PFAs elicited a concentration-dependent
inhibition of cellular communication in rat liver epithelial cells (Figure 18 and Table 14).
As consistent to previous results, inhibition potential of gap junction cellular
communication was dependent on the chain length of PFAs irrespective of its head group
(Upham et al., 1988). Only PFAs with a carbon-length of 5 to 9 in their fluorinated
backbone block the cellular communication. Among those potent PFAs, four PFAs
(PFOS, PFHS, PFOA, and PFDA) were detected in egg yolks. PFOS (EC50 = l.3><10l
mg/L) was the most potent among the PFAs investigated except 6:2 FTOH which was not
present in the sample and the potency decreased in the order of PFHS (1.7><10l mg/L),
PFOA (2.2><10l mg/L), and PFDA (2.3><10l mg/L). Previous studies have reported an
EC50 for PFOS of 1.5>< 101 mg/L (Hu et al., 2002). This observation demonstrates both

the robust nature of the assay and the consistency of results.

108

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A TEF value for each PFA was obtained by normalizing the EC50 concentrations
to PFOS EC50 (TEFpFA = ‘ECSOPFOS / ECSOPFA) (Table 14), which showed TEF values of
7.5x10", 5.6><10", and 4.7><10'l for PFHS, PFOA, and PFDA, respectively. A PFOS-
equivalent concentration (PEQ) was calculated for each mixture of PFAs by multiplying
TEFPFA with corresponding concentration in egg samples and summing (Villeneuve et al.,
2000). It should be noted that the present relative toxic potencies were derived from a
mammalian cell-bioassay, since methods for avian cells are not currently developed.
However, this cell-bioassay allows the comparison of toxicities of individual PFAs.
Furthermore, the relative potency of PFAs to alter GJIC is correlated with other
biological endpoints (Trosko and Ruch, 1998; Trosko et al., 1998). Therefore, the results
of this epigenetic test can be considered as being predictive of potential toxicity of PFAs

in mixture.

110

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Table 14. EC50 of PFAs and toxicity equivalent factor (TEF). PFOS equivalent
concentrations were calculated by multiplying TEF and individual PFA concentration in
a sample.

 

PFA C-lengths EC50 TEF
PFBA 3 (4) N.A. N.A.
PFBS 4 (4) 53.38 0.23
PFHA 5 (6) 31.15 0.40
PFHS 6 (6) 16.56 0.75
6:2 FTOH 6( 8) 11.29 1.11
PFOA 7 (8) 22.31 0.56
PFOS 8 (8) 12.48 1.00
8:2 FTOH 8(10) 11.00 1.13
PFDA 9 (10) 26.30 0.47
10:2 FTOH 10 (12) N.A. N.A.
10:1 FTOH 10 (11) N.A. N.A.
PFTeDA l3 (l4) N.A. N.A.

 

Risk Characterization of PF OS and other PFAs

Ecological risks of PF OS and PFAs to birds in the Lake Shihwa region were evaluated by
using two approaches. First, concentrations PFOS or the PF OS-EQs in eggs were
compared with toxicological benchmarks that represent thresholds below which adverse
effects on birds would not be expected. These benchmark values for avian species were
determined using the most ecologically relevant endpoints with uncertainty factors
assigned so that they are protective (Beach et al., 2006). Second, an average daily intake
(ADI) value was determined for PFOS based on the concentrations of PF As in the diet of
birds. The ADI benchmark dose was then compared with the calculated dietary dose. To

estimate the risk associated with the protection of 90% of birds in a population inhabiting

112

the region surrounding Lake Shihwa, a cumulative probability function was developed
for PFOS concentrations in egg yolks of the three species examined (Figure 19).

Hazard quotients (HQ = sample concentration/benchmark dose) were calculated
to provide preliminary estimates of risks associated with PFOS concentrations in birds.
Two toxicological benchmarks for PFOS in liver, serum, and egg yolk were reported
from dietary exposure studies with mallard and bobwhite quail (Newsted et al., 2005).
The lowest observable adverse effect level (LOAEL) and toxicity reference value (TRV)
for birds were determined to be 6.2><104 ng PFOS/mL and 1.7><103 ng PFOS/mL in egg
yolk, respectively. In the present study, the 90th centile of PFOS concentrations in the
yolk of bird eggs from the Lake Shihwa area was 4.8><102 ng/ g ww and the corresponding
HQ was 8.0><10'3 based on the LOAEL and 2.8><10'1 based on the TRV. As discussed
earlier, concentrations of PFCAs measured in eggs from the Lake Shihwa area were
greater than those in most other areas. Therefore, PFOS-EQs, based on relative potencies
in the GJIC assay, were calculated to assess the risk of PFAs mixture in egg yolk. In this
estimation, a slightly greater 90th centile (5.4><102 ng PFOS-EQ/g ww) resulted in greater

HQs (LOAEL = 9.0><10'3 and TRV = 3.2><10") but none exceeded a value of 1.0.

113

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Risk from dietary exposure to PFOS was evaluated for the population of little
egrets. Little egrets feed on various aquatic organisms, but during egg sampling, only
mullet carcass were observed around the little egret colony. Thus, in this diet-based
assessment, mullet caught in shallow waters of Lake Shihwa were assumed to be the sole
diet of the little egret. Toxicological doses for PFOS for a generic avian trophic level IV
predator were used as the benchmark for these calculations (LOAEL = 7.7><10'I mg
PFOS/kg/d and TRV = 22>< 10'2 mg PFOS/kg/d) (Newsted et al., 2005). The amount of
food ingested (F1) per day to body weight for wading birds was calculated for the little
egret (5x102 - 5x 103g ww) (FI = 9.7x 10" log (BW) —6.4><10", BW= body weight of bird)
(Kushlan et al., 1978). The mean concentrations of PFOS in carcass of mullet were
3.9><10I i 6.6 ng/g ww (Yoo et al., submitted). The ADI determined for little egret was
7.1 ><10'3 mg PFOS/kg/d. Based on this ADI value, HQs (=ADI/benchmark dose) were
calculated to range from 8><lO'3 - 9><10'3 based on the LOAEL and 3.l><lO'l - 3.3><10'l
based on the TRV.

Estimated risks of PFAs in avian species based on residue concentration in egg
yolk (upper 10% of bird population) and based on dietary exposure approaches were
quite similar. Although some assumptions have been made with limited toxicological
data, the comparable HQs derived allow us to evaluate the risks associated with current
exposures from PFOS and PFAs mixtures in Lake Shihwa at the screening level.
Approximately hundred-fold difference between PFAs exposure level and the threshold
values suggests that immediate threats such as reproductive failure are unlikely to occur
due to PFAs in birds in the Lake Shihwa region. In addition, considering the

conservative nature of TRV values (Beach et al., 2006), current concentrations of PFOS

115

and the mixture of PFAs would not be expected to pose adverse effects to the avian
population residing around Lake Shihwa. Nevertheless, further monitoring studies on the
health of bird populations and sources of PFAs in this region are warranted as the new

data pertaining to toxicities of PFAs become available.

116

CONCLUSION

As a first risk assessment for PFAs in wildlife to date in Korea, overall results
suggest a low ecological risk of PFAs exposure to marine wildlife in the Lake Shihwa
area. In general, contamination status of PFAs in biota is comparable to other urban
areas. However, as mentioned earlier, most of toxicological benchmarks for wildlife
have been derived using a limited number of data generated with few endpoints to
evaluate the health of aquatic wildlife. Toxic effects from chronic PFA exposure are not
well studied at present. Therefore, more ecologically meaningful benchmark values are
needed to draw conclusions about the status of environmental health. Finally, regional
monitoring programs should be established in this area until we are confident that on-
going world-wide efforts to reduce the emission of PFA into the environment are

working as well as intended.

ll7

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