Arch. Environ. Contam. Toxicol. 38, 362–370 (2000)
DOI: 10.1007/s002449910048
A R C H I V E S O F
Environmental
Contamination
a n d Toxicology
r 2000 Springer-Verlag New York Inc.
Polychlorinated Dibenzo-p-Dioxins (PCDDs), Dibenzofurans (PCDFs),
Biphenyls (PCBs), and Organochlorine Pesticides in Yellow-Blotched
Map Turtle from the Pascagoula River Basin, Mississippi, USA
K. Kannan,1 M. Ueda,2 J. A. Shelby,3 M. T. Mendonca,3 M. Kawano,2 M. Matsuda,2 T. Wakimoto,2 J. P. Giesy1
1
213 National Food Safety and Toxicology Center, Department of Zoology, Institute for Environmental Toxicology, Michigan State University,
East Lansing, Michigan 48824, USA
2 Department of Environment Conservation, Ehime University, Tarumi 3-5-7, Matsuyama 790-8566, Japan
3 Department of Zoology and Wildlife, Auburn University, Auburn, Alabama 36849, USA
Received: 4 May 1999/Accepted: 20 September 1999
Abstract. Concentrations of polychlorinated-dibenzo-p-dioxins (PCDDs), -dibenzofurans (PCDFs), -biphenyls (PCBs), and
organochlorine pesticides were measured in tissues of map
turtles collected from two locations in the Pascagoula River
drainage of Mississippi, USA. PCBs were most predominant
among the organochlorines with a concentration of up to 99
ng/g, wet weight (580 ng/g, lipid weight) in livers. The greatest
concentration of PCDDs/DFs of 1.1 pg/g, wet weight (15.76
pg/g, lipid weight) was found in the liver of a male turtle. The
measured concentrations of organochlorines were less than
those reported for turtles from the Great Lakes Basin and upper
St. Lawrence River. PCBs contributed 90–99% of the total estimated
2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents (TEQs). Particularly, PCB congeners 105, 118, and 156 accounted for 68–80% of
the estimated toxic potency of PCBs in turtles.
The yellow-blotched map turtle (sawback), Graptemys flavimaculata Cagle, is found only in the Pascagoula River system
of southern Mississippi, USA. It is known from the Pascagoula
River and its major tributaries, Leaf, Chickasawhay, and
Escatawpa Rivers. Due to the rapid decline in the populations,
this species is classified as threatened under the Endangered
Species Act by the U.S. Fish and Wildlife Service (Seigel and
Brauman 1995). The reason for this decline is not known, but
high rates of mortality in nests from river flooding, nest
predation, and water quality degradation have been suggested
as possible causes (Jones 1992; Seigel and Brauman 1995).
Reproductive frequency, clutch size, and hatching success of
yellow-blotched map turtles were less than those of the other
related turtles in Mississippi (Seigel and Brauman 1995).
Persistent, bioaccumulative, and toxic contaminants, such as
DDTs and polychlorinated biphenyls (PCBs), have been suspected to cause reproductive impairment in certain reptilians,
including some turtle species, because of vulnerability of early
Correspondence to: K. Kannan
life stages (Bergeron et al. 1994; Guillette et al. 1995; Bishop et
al. 1998). Red-eared slider turtle eggs (Tachyemys scripta)
exposed to certain PCB congeners and their hydroxylated
metabolites produced females at male-producing temperatures
(Bergeron et al. 1994). Snapping turtles (Chelydra serpentina
serpentina) that contained greater concentrations of DDTs and
PCBs in eggs had higher rates of abnormalities (poor development and hatching success) than those that contained relatively
less concentrations (Bishop et al. 1991). Adult snapping turtles
from the Hudson River, New York, have been reported to
accumulate several parts per million (ppm) concentrations of
PCBs in tissues (Olafsson et al. 1983; Bryan et al. 1987a,
1987b; de Solla et al. 1998). Reports of organochlorine
accumulation in yellow-blotched map turtles were not available. In this study, concentrations of PCB congeners, organochlorine (OC) pesticides, polychlorinated dibenzo-p-dioxins (PCDDs), and polychlorinated dibenzofurans (PCDFs) were
measured in liver, muscle, and fat of yellow-blotched map
turtles collected from two locations in the Pascagoula River
drainage in Mississippi. 2,3,7,8-Tetrachlorodibenzo-p-dioxin
equivalents (TEQs) of PCBs, PCDDs, and PCDFs were estimated to evaluate their relative contribution to toxicity.
Materials and Methods
Yellow-blotched map turtles were collected under letters of permission
from Mississippi State Department of Wildlife, Fisheries and Parks in
Hattiesburg and Vancleave, Mississippi (Figure 1), areas that still
support relatively large populations (e.g., 2,000 individuals). Turtles
were caught in the Pascagoula River basin near Hattiesburg and
Vancleave. Turtles were trapped in baskets attached to logs where they
were frequently seen basking. Upon capture, turtles were sexed,
transferred to the laboratory, and euthanized with Nembutal (45
mg/kg). Animals were frozen at 0°C after sacrifice. Animals were
partially thawed and dissected at a later date to collect liver, muscle,
and fat tissues. Samples were stored in chemically cleaned jars at
⫺20°C until analysis. Muscle from five animals, liver from six
animals, and fat from three animals were taken. Due to the availability
of a relatively small mass of fat, samples were pooled for analysis. OC
pesticides and PCBs were analyzed in livers of four animals and
PCDDs, PCDFs, PCBs, and OC Pesticides in Map Turtles
363
Fig. 1. Map of Mississippi, USA, showing
Pascagoula River basin and the sampling
locations of yellow-blotched map turtles
muscle from a single animal, whereas PCDDs/DFs were analyzed in
five livers, four muscle, and one fat samples.
Chemical Analysis
OC pesticides, PCBs, PCDDs, and PCDFs were analyzed following
methods described in detail elsewhere (Nakamura et al. 1993; Kannan
et al. 1995; Khim et al. 1999), with some modifications. Sample tissues
(⬃30 g for muscle and liver, ⬃3–6 g for fat) were homogenized with
anhydrous sodium sulfate and Soxhlet extracted with methylene chloride
(DCM) and hexane (3:1; 400 ml) for 20 h. Extracts were concentrated using
a rotary evaporator at 40°C, and lipid content was measured from an
aliquot of the extract by gravimetry. Sample extracts were split to
determine PCDDs/PCDFs and the remaining OCs separately.
For PCB and OC pesticides analyses, sample extracts were treated
with sulfuric acid, passed through activated florisil (10 g; 60–100 mesh
size; Sigma, St. Louis, MO) packed in a glass column (10 mm ID) for
cleanup and fractionation. The first fraction eluted with 100 ml of high
purity hexane (Burdick and Jackson, Muskegon, MI) contained PCBs,
HCB and p,p8-DDE. Remaining OC pesticides were eluted in the
second fraction using 100 ml 20% DCM in hexane.
OC pesticides such as DDTs ( p,p8-DDE, p,p8-DDD, p,p8-DDT),
HCHs (␣-, ␤-, ␥-isomers), CHLs (␣-chlordane, ␥-chlordane, heptachlor epoxide), and HCB and PCBs were quantified using a gas
chromatograph (Perkin Elmer series 600) equipped with 63Ni electron
capture detector (GC-ECD). A fused silica capillary column coated
with DB-5MS [(5%-phenyl)-methylpolysiloxane, 30 m ⫻ 0.25 mm
ID; J&W Scientific, Folsom, CA] having a film thickness of 0.25 µm
was used. The column oven temperature was programmed from 120°C
(1 min hold) to 180°C at a rate of 10°C/min (1 min hold) and then to
260°C at a rate of 2°C/min with a final holding time of 12 min. Injector
and detector temperatures were kept at 250°C and 300°C, respectively.
Helium was used as the carrier gas and nitrogen was the make-up gas.
An equivalent mixture of Aroclors 1016, 1242, 1254, and 1260 with
known composition and content was used as a standard. Individual
PCB congeners were identified using a standard mixture containing
100 PCBs. Concentrations of individually resolved peaks were summed
to obtain total PCB concentrations. PCB congeners are referred by
Ballschmiter and Zell numbers throughout the manuscript. OC pesticides were quantified from individually resolved peak areas based on
the peak areas of standards. Recoveries of PCB congeners and OC
pesticides spiked into samples and passed through the analytical
procedure were between 85 and 105%. Detection limit for OC
pesticides was 0.01 ng/g and for all PCBs was 1.0 ng/g, wet weight.
364
K. Kannan et al.
Table 1. Concentrations of organochlorine pesticides and polychlorinated biphenyls (ng/g, wet weight) in tissues of female yellow-blotched
map turtles
Sample No.
106
810-8911
811-8912
211
39
Location
Tissue
Fat (%)
Total PCBs
DDTsa
HCHsb
CHLsc
HCB
Hattiesburg
Liver
17
99 (580)
50 (290)
1.0 (5.9)
43 (250)
14 (82)
Vancleave
Liver
25
32 (130)
7.7 (31)
0.17 (0.68)
3.1 (12)
1.5 (6)
Vancleave
Liver
26
33 (130)
11 (42)
0.71 (2.7)
3.2 (12)
1.8 (6.9)
Vancleave
Liver
4.8
15
(310)
4.0 (83)
0.49 (10)
3.7 (77)
NDd
Vancleave
Muscle
2.2
7.1 (320)
⬍ 0.1 (⬍ 4.5)
⬍ 0.1 (⬍ 4.5)
⬍ 0.1 (⬍ 4.5)
⬍ 0.1 (⬍ 4.5)
DDTs ⫽ p,p8-DDE ⫹ p,p8-DDD ⫹ p,p8-DDT
HCHs ⫽ ␣ ⫹ ␤ ⫹ ␥ isomers
c CHLs ⫽ ␣-chlordane ⫹ ␥-chlordane ⫹ heptachlor epoxide
d ND ⫽ Not determined
a
b
Seventeen 13C-labeled tetra-, penta-, hexa-, hepta-, and octa-CDD
and CDF congeners substituted at the 2,3,7,8-positions (Cambridge
Isotope Laboratories Inc., MA), were spiked at 500 pg each into hexane
extracts prior to sulfuric acid treatment. Extracts were passed through a
silica gel packed glass column (Merck, silicagel 60; 3 g) and eluted
with 120 ml hexane. The hexane extract was K-D (Kuderna-Danish)
concentrated and passed through alumina column (Merck-Alumia
oxide, activity grade 1) and eluted with 200 ml hexane as the first
fraction, which was discarded. The second fraction eluted with 80 ml of
50% DCM in hexane was K-D concentrated and passed through silica
gel impregnated activated carbon column (1 g) (Tsuda et al. 1993). The first
fraction eluted with 60 ml 25% DCM in hexane was discarded, and the
second fraction eluted with 200 ml toluene was concentrated and analyzed
by a high-resolution gas chromatograph interfaced with a high-resolution
mass spectrometer (HRGC-HRMS).
A Hewlett Packard 5890 series II HRGC connected to a JEOL JMS
SX102A HRMS was used for the identification and quantification of
individual congeners. Separation of tetra-, penta-, and hexa-chlorinated
dibenzo-p-dioxin (CDD) and chlorinated dibenzofuran (CDF) was
achieved by a CP-Sil 88 column (Chrompack; 60 m ⫻ 0.25 mm ID)
coated at 0.25 µm film thickness, whereas those of hepta- and
octa-CDD and CDF was achieved by a DB-5 column (J&W Scientific;
30 m ⫻ 0.25 mm ID) coated at 0.2 µm film thickness. The column oven
temperature for both CP-Sil and DB-5 columns were held at 100°C for
1 min, and then increased at a rate of 20°C/min to 180°C, held for 1 min, and
then increased at a rate of 4°C/min to 240°C. The injection was made
splitless. Both injector and transfer line temperatures were held at 250°C.
Helium was used as a carrier gas at a flow rate of 1 ml/min. The mass
spectrometer conditions were as follows: Ionization mode, EI; electron
energy, 70eV. PCDD and PCDF congeners were determined by selected ion
monitoring (SIM) at the two most intensive ions of molecular ion cluster.
Recoveries of individual PCDD and PCDF congeners varied from 33 to
114%, depending on the sample and congeners. The detection limits of
tetra-, penta-, hexa-, hepta-, and octa-CDDs and CDFs were 0.008, 0.01,
0.01, 0.02, 0.02 pg/g, lipid weight, respectively. Recovery of internal
standards (13C-labeled PCDDs/DFs), on average was, 56%. Reported
concentrations were not corrected for the recoveries of internal standard.
Procedural blanks were analyzed through the whole procedure to check for
interferences and contamination arising from glassware and solvents.
Results and Discussion
PCBs and OC Pesticides
PCBs were the predominant group of OC compounds found in
turtle tissues (Table 1). Concentrations of OC pesticides were
less than those of PCBs and were found in the order DDTs ⬎
CHLs ⬎ HCB ⬎ HCHs. p,p8-DDE and ␤-HCH accounted for
greater than 70% of the corresponding total concentrations of
DDTs and HCHs. This suggested that these compounds were
derived from historical inputs rather than recent inputs. On a
wet weight basis, OC concentrations in livers were greater than
that in muscle, which could be explained by the differences in
the lipid content between liver and muscle.
The lipid-normalized concentrations of PCBs in livers varied
from 130 to 580 ng/g. Variations in the concentrations of PCBs
among individuals may have been related to their age and size.
Although age and size of the turtles were not measured in this
study, an earlier study showed that the concentrations of OCs in
livers of snapping turtles varied with age and size (Hebert et al.
1993). Hepatic concentration of PCBs in turtle from Hattiesburg was two to five times greater than that from Vancleave.
However, the number of samples analyzed is small to provide
conclusive evidence on the spatial differences in concentrations.
No earlier studies have reported the concentrations of PCBs
and OC pesticides in yellow-blotched map turtles. Therefore,
concentrations measured in map turtle tissues were compared
with those reported for snapping turtles, for which considerable
information is available (Meyers-Schöne and Walton 1994).
Concentrations of PCBs in livers of map turtles were approximately 1,000 times less than those reported for snapping turtle
livers from a contaminated site in the Upper Hudson River,
New York (Bryan et al. 1987a), and 10 to 100 times less than
those of diamondback terrapins (Malaclemys terrapin) from a
PCB-contaminated Superfund site in Georgia (Kannan et al.
1998). Nevertheless, total PCB concentrations in map turtle
livers were similar to those reported in the flesh of snapping
turtles from lakes and rivers in Minnesota (Helwig and Hora
1983). Relatively less accumulation of organochlorines in
yellow-blotched map turtles than in snapping turtles may be
explained by their food habits. While the diet of snapping
turtles consists of fish (33%), amphibians, crustaceans, birds
and bird eggs (33%), and plant materials (33%) (Bishop et al.
1994), map turtles feed on insects (50–60%), sponges (35–50%),
algae (0–30%), and mollusks (Seigel and Brauman 1994).
The pattern of relative concentrations of PCB congeners in
turtle tissues was dominated by relatively few higherchlorinated congeners (Figure 2). Hexa- and hepta-chlorobiphe-
PCDDs, PCDFs, PCBs, and OC Pesticides in Map Turtles
365
Fig. 2. Relative abundance of PCB isomers and
congeners in livers of female turtles collected from
Hattiesburg and Vancleave in Pascagoula River Basin. Abundances of individual PCB congener is normalized to the PCB congener 153 (2,28,4,48,5,58-),
which is treated as 100. Abundances are in
weight %
nyls accounted for 60–71% of the total PCB concentrations
(Figure 3). Turtles from Vancleave contained a greater proportion of higher-chlorinated PCB congeners than the one from
Hattiesburg. This suggests differences in the sources of PCB
exposure to turtles between the two sampling locations. Hexachlorobiphenyl 153 (2,28, 4,48,5,58) was the predominant
congener in turtle liver, accounting for 14 (Hattiesburg) to 21%
(Vancleave) of the total PCB concentrations (Figure 2). PCB
congeners 153, 180 (2,28,3,4,48,5,58), 118 (2,38,4,48,5), and 138
(2,28,3,4,48,58) collectively accounted for 45–50% of the total
PCB concentrations in turtle livers. Muscle tissue of a turtle
from Vancleave also contained a similar congener distribution
to that in liver except for the presence of a greater proportion of
congeners 180 (14%) and 170 (13%) than that of 153 (12%) in
the total PCB concentrations. These results suggest that turtles
were exposed to highly chlorinated PCB mixtures, such as
Aroclors 1254 and 1260, or that they metabolize lesschlorinated PCB congeners efficiently (Yawetz et al. 1997).
The chlorobiphenyl congeners exhibiting the greatest toxicity
toward mammals are isostereomers of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and include non-ortho coplanar PCB
congeners 77, 126, and 169 (Tanabe 1988). Mono-orthosubstituted PCB congeners, such as PCB 118, 105, 156, also
exhibit similar but weaker toxic responses (Van den Berg et al.
1998). Since these congeners elicit common toxic responses
through a common mode of action (Giesy and Kannan 1998),
their concentrations are expressed as toxic equivalents to
identify those congeners that may pose greater health risks.
Toxic equivalents (TEQs) were calculated by multiplying the
concentration of each of the toxic congeners by its toxicity
equivalency factor (TEF). TEFs for PCB congeners have been
derived based on their toxicity relative to TCDD, which is
considered the most potent congener.
The non-ortho PCB congeners 77, 126, and 169 were not
found in turtle livers from Vancleave (Table 2). However, livers
of turtles collected from Hattiesburg contained PCB congener
77 at a mean concentration of 100 pg/g, wet weight. Mono- and
di-ortho congeners were found in all samples from both
Vancleave and Hattiesburg. The estimated TEQs of PCBs in
turtle livers from Hattiesburg and Vancleave were 4.4 and 0.71
(range: 0.33–0.93) pg/g, wet weight, respectively. When TEQs
for congeners 126 and 169 were considered at the limits of
Fig. 3. Homolog composition of PCBs in the livers of map turtles from
Hattiesburg and Vancleave
detection, the upper limits for TEQs in turtles from Hattiesburg
and Vancleave were 9.8 and 6 pg/g, wet weight, respectively.
PCB congeners 156, 118, and 105 contributed 68–80% of the
total TEQs. Similarly, greater contributions of mono-ortho PCB
congeners 156, 118, and 105 to the TEQs have been reported in
marine mammals (Kannan et al. 1993; Corsolini et al. 1995).
These results indicate that congeners 156, 118, and 105 would
potentially account for most toxicity to the health of these
turtles.
PCDDs and PCDFs
Concentrations of 17 2,3,7,8-substituted PCDD and PCDF
congeners in fat, muscle, and liver of turtles are shown (Table
3). Concentrations of PCDD/DFs in the liver of one female
turtle from Vancleave (2.0–6.3 pg/g, lipid weight) were three
times higher than those found in Hattiesburg (2.0 pg/g, lipid
weight). The greatest concentration of PCDDs was found in the
liver of a male turtle (15.8 pg/g, lipid weight), which was eight
times greater than that found in the livers of female turtles from
Hattiesburg. The concentrations of PCDDs and PCDFs were
less than those reported for the eggs of snapping turtles from the
Great Lakes basin (Bishop et al. 1996) and the upper St.
Lawrence River (Ryan et al. 1986).
366
K. Kannan et al.
Table 2. Concentrations of non-, mono-, and di-ortho coplanar PCBs (pg/g wet weight) and their 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents
(TEQs) in livers of female yellow-blotched map turtles
PCB
Congener
Hattiesburg (n ⫽ 1)
Structure
TEFb
3,38,4,483,38,4,48,53,38,4,48,5,58-
0.0001
0.1
0.01
2,3,38,4,48-
118
TEQ
Conc
TEQ
100
⬍ 50
⬍ 40
0.01
⬍5
⬍ 0.4
⬍ 40
⬍ 50
⬍ 40
⬍ 0.004
⬍5
⬍ 0.4
0.0001
3,860
0.39
2,38,4,48,5-
0.0001
12,600
1.26
156
2,3,38,4,48,5-
0.0005
3,700
1.85
157
2,3,38,4,48,58-
0.0005
620
0.31
167a
2,38,4,48,5,58-
0.00001
2,000
0.02
580
(350–700)c
1,900
(810–2,600)
460
(230–590)
140
(20–210)
740
(360–1,000)
0.06
(0.035–0.07)
0.19
(0.08–0.26)
0.23
(0.11–0.29)
0.07
(0.01–0.1)
0.01
(0.004–0.01)
Di-ortho
170
2,28,3,38,4,48,5-
0.0001
4,500
0.45
180
2,28,3,4,48,5,58-
0.00001
11,000
0.11
1,300
(690–1,600)
3,300
(1,700–4,300)
0.13
(0.07–0.16)
0.03
(0.02–0.04)
0.71
(0.33–0.93)
Non-ortho
77
126
169
Mono-ortho
105
Conc
Vancleave (n ⫽ 3)
Total TEQs
4.40
a
Coelutes with congener 185
From Van den Berg et al. (1998)
c Values in parentheses indicate range
b
Concentrations of PCDDs were greater than that of PCDFs in
the liver (Table 3). Concentrations of PCDDs in turtle livers
were 7 to 75 times greater than those of PCDFs. Concentrations
of PCDFs in turtle muscle were below the limit of detection in
all the samples except one, which contained 2,3,7,8-TCDF at
0.40 pg/g, lipid weight.
Among various PCDD congeners, OCDD accounted for the
greatest proportion of total PCDD concentrations (Figure 4).
The compositions of individual PCDD congener to the concentrations of total PCDDs in livers of female turtles from
Hattiesburg and Vancleave were different (Figure 4). A higher
proportion of OCDD to the total PCDD concentration was
found in turtles from Vancleave (68%) than those from Hattiesburg (43%). Nevertheless, the composition of PCDF congeners
in the livers of turtles from Vancleave and Hattiesburg was
similar.
Estimated concentrations of TEQs contributed by PCDD/DF
congeners in the livers of turtles from Vancleave and Hattiesburg were 0.414 and 0.219 pg/g, lipid weight, respectively
(Table 4). PCDDs contributed greater than 90% of the PCDD/DF
TEQs in turtle livers. 2,3,7,8-TCDD accounted for 51% of the
total TEQs contributed by PCDDs in turtles from Vancleave.
The contribution by 2,3,7,8-TCDD to the total TEQs in turtles
from Hattiesburg was 29%, which was less than that found in
Vancleave. Among PCDF congeners, 1,2,3,7,8-PeCDF contributed to 50% of the TEQs. Despite the presence of great
concentrations of OCDD in turtle livers, its contribution to
TEQs was insignificant (Figure 4).
The lipid-normalized concentrations of TEQs estimated for
PCBs in turtle liver from Vancleave and Hattiesburg were 3.8
and 26 pg/g, lipid weight, respectively. Generally, TEQs
contributed by PCBs in turtles from Hattiesburg and Vancleave
were approximately 10 and 100 times greater than those
contributed by PCDDs and PCDFs. Overall, PCBs contributed
to 90–99% of the total TEQ concentrations in female turtle
livers (Figure 5). The contribution by PCDDs to the total TEQs
in turtles was 9.5% in Vancleave samples and 0.8% in
Hattiesburg samples. These results suggest greater exposure of
turtles collected in Vancleave to PCDD sources than those from
Hattiesburg.
Health Implications of Turtles
The toxicological significance of OC concentrations measured
in turtle tissues is difficult to assess because of the lack of
toxicity reference doses. Earlier studies have measured OC
concentrations in the eggs of turtles, since the early life stages
are more susceptible to the toxic effects (Hebert et al. 1993;
Bishop et al. 1994, 1996). Studies have suggested a significant
correlation between the concentrations of OCs in turtle livers
and eggs, indicating that an examination of contaminant levels
in liver tissues might also provide information regarding the
contaminants in egg (Hebert et al. 1993). Exposure of green sea
turtle (Chelonia mydas) eggs to p,p8-DDE at concentrations of
up to 543 ng/g wet weight, did not alter sex ratio and hatching
success (Podreka et al. 1998). Based on the assumption that
lipid-normalized concentrations of OCs in turtle eggs would be
similar to that found in the liver (Meyers-Schöne and Walton
1994), the measured concentrations of p,p8-DDE in map turtles
were less than that would affect hatching success and survival
rate. Female turtles fast during the nesting season, which may
PCDDs, PCDFs, PCBs, and OC Pesticides in Map Turtles
367
Table 3. Concentrations of PCDD/PCDF congeners (pg/g, lipid weight) in tissues of yellow-blotched map turtles
Sample
Location
Sex
Tissue
Fat (%)
PCDDs
2,3,7,8-TCDD
1,2,3,7,8-PeCDD
1,2,3,4,7,8-HxCDD
1,2,3,6,7,8-HxCDD
1,2,3,7,8,9-HxCDD
1,2,3,4,6,7,8-HpCDD
OCDD
Total PCDD
Surrogate 1 recovery (%)a
PCDFs
2,3,7,8-TCDF
1,2,3,7,8-PeCDF
2,3,4,7,8-PeCDF
1,2,3,4,7,8-HxCDF
1,2,3,6,7,8-HxCDF
1,2,3,7,8,9-HxCDF
2,3,4,6,7,8-HxCDF
1,2,3,4,6,7,8-HpCDF
1,2,3,4,7,8,9-HpCDF
OCDF
Total PCDF
Surrogate 2 recovery (%)b
Total PCDD/PCDF
a
b
Pool
Pool
Pool
106 & 110 106 & 110 64 & 65
Pool 89-8912,
811-8912,
810-8911
89-8912
89-8912
810-8911 810-8911 811-8912 811-8912
Hattiesburg Hattiesburg Hattiesburg Vancleave
Female
Female
Male
Female
Muscle
Liver
Liver
Fat
1.4
24
7
70
Vancleave Vancleave Vancleave Vancleave Vancleave Vancleave
Female
Female
Female
Female
Female
Female
Muscle
Liver
Muscle
Liver
Muscle
Liver
1.6
13
1.3
17
2.3
19
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
1.14
1.139
40
0.061
0.103
0.035
0.310
0.066
0.503
0.818
1.894
35
⬍ 0.008
0.662
⬍ 0.01
1.93
⬍ 0.01
1.4
11.6
15.592
35
0.020
0.010
0.018
0.040
0.013
0.059
0.516
0.677
63
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
0.99
0.990
41
0.255
0.197
0.082
0.572
0.091
0.331
4.66
6.188
48
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
0.828
0.828
37
0.177
0.089
0.059
0.220
0.063
0.237
1.477
2.322
39
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
0.338
0.338
51
0.164
0.16
0.062
0.257
0.066
0.233
0.792
1.734
36
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.020
0.021
0.007
0.023
⬍ 0.01
⬍ 0.01
⬍ 0.01
0.037
⬍ 0.02
⬍ 0.02
0.107
97
2.001
0.169
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.169
78
15.761
0.009
⬍ 0.01
⬍ 0.01
⬍0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.009
64
0.686
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.037
0.046
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.083
42
6.271
0.396
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.396
87
1.224
0.079
0.039
0.023
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.142
60
2.464
⬍ 0.008
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.01
⬍ 0.02
⬍ 0.02
⬍ 0.02
0.04
0.082
0.009
0.041
⬍ 0.01
⬍ 0.01
⬍ 0.01
0.069
⬍ 0.02
⬍ 0.02
0.241
59
1.975
52
1.139
54
0.990
114
0.338
Surrogate 1 ⫽ 13C TCDD and 13C OCDD, respectively
Surrogate 2 ⫽ 13C TCDF and 13C OCDF, respectively
Fig. 4. Composition of individual PCDD and PCDF
congeners to the corresponding total PCDD and
PCDF concentrations (top) and TEQs (bottom) in the
livers of female map turtles from Vancleave and Hattiesburg in the Pascagoula River basin
result in the mobilization of stored lipids and consequently OCs
to eggs (Meyers-Schöne and Walton 1994). PCB concentrations
in eggs were 10 to 64 times greater than those in the muscle of
snapping turtles (Bryan et al. 1987b). The use of fat reserves
results in the mobilization and preferential deposition of PCBs
into the lipid-rich egg yolks. A few field studies from the Great
Lakes–St. Lawrence River Basin have correlated greater concentrations of PCBs and other OCs in turtle eggs to poor hatching
success and deformities (Bishop et al. 1991, 1996; de Solla et
al. 1998). Therefore, measurement of OC concentrations in
a
0.0612
0.1025
0.0035
0.0310
0.0066
0.0050
0.0001
0.2098
0.0020
0.0010
0.0036
0.0023
0.0004
0.0092
0.2190
⬍ 0.008
⬍ 0.01
⬍ 0.001
⬍ 0.001
⬍ 0.001
⬍ 0.0002
0.0001
0.0001
⬍ 0.0008
⬍ 0.0005
⬍ 0.005
⬍ 0.001
⬍ 0.0002
1
1
0.1
0.1
0.1
0.01
0.0001
0.1
0.05
0.5
0.1
0.01
0.0001
Hattiesburg
Female
Liver
24
Pool
106 & 110
Hattiesburg
Female
Muscle
1.4
TEFa
From Van den Berg et al. (1998)
Location
Sex
Tissue
Fat (%)
PCDDs
2,3,7,8-TCDD
1,2,3,7,8-PeCDD
1,2,3,4,7,8-HxCDD
1,2,3,6,7,8-HxCDD
1,2,3,7,8,9-HxCDD
1,2,3,4,6,7,8-HpCDD
OCDD
Total PCDD TEQs
PCDFs
2,3,7,8-TCDF
1,2,3,7,8-PeCDF
2,3,4,7,8-PeCDF
1,2,3,4,7,8-HxCDF
1,2,3,4,6,7,8-HpCDF
Total PCDF TEQs
Total PCDD/PCDF TEQs
Sample
Pool
106 & 110
0.0169
⬍ 0.0005
⬍ 0.005
⬍ 0.001
⬍ 0.0002
0.0169
0.8871
⬍ 0.008
0.662
⬍ 0.001
0.193
⬍ 0.001
0.014
0.0012
0.8702
Hattiesburg
Male
Liver
7
Pool
64 & 65
0.0009
⬍ 0.0005
⬍ 0.005
⬍ 0.001
⬍ 0.0002
0.0009
0.0392
0.0204
0.0101
0.0018
0.0040
0.0013
0.0006
0.0001
0.0383
Vancleave
Female
Fat
70
Pool 89-8912,
811-8912,
810-8911
0.0001
⬍ 0.0008
⬍ 0.0005
⬍ 0.005
⬍ 0.001
⬍ 0.0002
⬍ 0.008
⬍ 0.01
⬍ 0.001
⬍ 0.001
⬍ 0.001
⬍ 0.0002
0.0001
0.0001
Vancleave
Female
Muscle
1.6
89-8912
0.0037
0.0023
⬍ 0.005
⬍ 0.001
⬍ 0.0002
0.006
0.5363
0.255
0.197
0.008
0.057
0.009
0.003
0.0005
0.5303
Vancleave
Female
Liver
13
89-8912
0.0396
⬍ 0.0005
⬍ 0.005
⬍ 0.001
⬍ 0.0002
0.0396
0.0397
⬍ 0.008
⬍ 0.01
⬍ 0.001
⬍ 0.001
⬍ 0.001
⬍ 0.0002
0.0001
0.0001
Vancleave
Female
Muscle
1.3
810-8911
Table 4. 2,3,7,8-Tetrachlorodibenzo-p-dioxin equivalents (TEQs) of PCDDs and PCDFs (pg/g, lipid weight) in yellow-blotched map turtle tissues
0.0079
0.0019
0.0116
⬍ 0.001
⬍ 0.0002
0.0215
0.3245
0.1769
0.0894
0.0059
0.0220
0.0063
0.0024
0.0001
0.3030
Vancleave
Female
Liver
17
810-8911
0.00003
⬍ 0.0008
⬍ 0.0005
⬍ 0.005
⬍ 0.001
⬍ 0.0002
⬍ 0.008
⬍ 0.01
⬍ 0.001
⬍ 0.001
⬍ 0.001
⬍ 0.0002
0.00003
0.00003
Vancleave
Female
Muscle
2.3
811-8912
0.0040
0.0041
0.0045
0.0041
0.0007
0.0174
0.3823
0.164
0.16
0.0062
0.0257
0.0066
0.0023
0.0001
0.3649
Vancleave
Female
Liver
19
811-8912
368
K. Kannan et al.
PCDDs, PCDFs, PCBs, and OC Pesticides in Map Turtles
Fig. 5. Contribution (%) by PCBs, PCDDs, and PCDFs to the total
TEQs in livers of map turtles from Vancleave and Hattiesburg in the
Pascagoula River basin (contribution by PCDFs to the TEQs was
insignificant)
eggs of yellow-blotched map turtles is needed. The concentrations of OCs measured in tissues of map turtles were lower than
those found in field studies that reported poor hatching success
and survival rates in snapping turtles. However, a recent study
has suggested that the exposure of eggs of red-eared slider turtle
(Trachemys scripta elegans) to OC pesticides and PCBs in
concert with one another produced significant sex reversal
compared to those that were incubated with a single OC
compound, suggesting the potential for interactive effects of
several OCs (Willingham and Crews 1999).
Acknowledgments. This research was supported by a grant from
National Institute of Health Superfund Basic Research Program
(NIH-ES-04911) to Michigan State University.
References
Bergeron JM, Crews D, McLachlan JA (1994) PCBs as environmental
estrogens: turtle sex determination as a biomarker of environmental contamination. Environ Health Perspect 102:780–781
Bishop CA, Brooks RJ, Carey JH, Ng P, Norstrom RJ, Lean DRS
(1991) The case for a cause-effect linkage between environmental
contamination and development in eggs of the common snapping
turtle (Chelydra s. serpentina) from Ontario, Canada. J Toxicol
Environ Health 33:521–547
Bishop CA, Brown GP, Brooks RJ, Lean DRS, Carey JH (1994)
Organochlorine contaminant concentrations in eggs and their
relationship to body size, and clutch characteristics of the female
common snapping turtle (Chelydra serpentina serpentina) in Lake
Ontario, Canada. Arch Environ Contam Toxicol 27:82–87
Bishop CA, Ng P, Norstrom RJ, Brooks RJ, Pettit KE (1996) Temporal
and geographic variation of organochlorine residues in eggs of the
common snapping turtle (Chelydra serpentina serpentina) (1981–
1991) and comparisons to trends in the herring gull (Larus
argentatus) in the Great Lakes basin in Ontario, Canada. Arch
Environ Contam Toxicol 31:512–524
Bishop CA, Ng P, Pettit KE, Kennedy S, Stegeman JJ, Norstrom RJ,
Brooks RJ (1998) Environmental contamination and developmental abnormalities in eggs and hatchlings of the common snapping
turtle (Chelydra serpentina serpentina) from the Great Lakes–St.
Lawrence basin (1989–91). Environ Pollut 99:1–14
Bryan AM, Olafsson PG, Stone WB (1987a) Disposition of low and
high environmental concentrations of PCBs in snapping turtle
tissues. Bull Environ Contam Toxicol 38:1000–1005
Bryan AM, Olafsson PG, Stone WB (1987b) Disposition of toxic PCB
congeners in snapping turtle eggs: expressed as toxic equivalents
of TCDD. Bull Environ Contam Toxicol 39:791–796
369
Corsolini S, Focardi S, Kannan K, Tanabe S, Borrell A, Tatsukawa R
(1995) Congener profile and toxicity assessment of polychlorinated biphenyls in dolphins, sharks and tuna collected from Italian
coastal waters. Mar Environ Res 40:33–53
de Solla SR, Bishop CA, van der Kraak G, Brooks RJ (1998) Impact of
organochlorine contamination on levels of sex hormones and
external morphology of common snapping turtles (Chelydra
serpentina serpentina) in Ontario, Canada. Environ Health Perspect 106:253–260
Guillette LJ Jr, Crain DA, Rooney AA, Pickford DB (1995) Organization versus activation: the role of endocrine-disrupting contaminants (EDCs) during embryonic development in wildlife. Environ
Health Perspect 103:157–164
Giesy JP, Kannan K (1998) Dioxin-like and non-dioxin-like toxic
effects of polychlorinated biphenyls (PCBs): implications for risk
assessment. Crit Rev Toxicol 28:511–569
Hebert CE, Glooschenko V, Haffner GD, Lazar R (1993) Organic
contaminants in snapping turtle (Chelydra serpentina) populations
from southern Ontario, Canada. Arch Environ Contam Toxicol
24:35–43
Helwig DD, Hora ME (1983) Polychlorinated biphenyl, mercury, and
cadmium concentrations in Minnesota snapping turtles. Bull
Environ Contam Toxicol 30:186–190
Jones RL (1992) Technical draft: yellow-blotched map turtle (Graptemys flavimaculata) recovery plan. US Fish and Wildlife Service,
Jackson, MS
Kannan K, Tanabe S, Borrell A, Aguilar A, Focardi S, Tatsukawa R
(1993) Isomer-specific analysis and toxic evaluation of polychlorinated biphenyls in striped dolphins affected by an epizootic in the
western Mediterranean Sea. Arch Environ Contam Toxicol 25:227–
233
Kannan K, Tanabe S, Tatsukawa R (1995) Geographical distribution
and accumulation features of organochlorine residues in fish in
tropical Asia and Oceania. Environ Sci Technol 29:2673–2683
Kannan K, Nakata H, Stafford R, Masson GR, Tanabe S, Giesy JP
(1998) Bioaccumulation and toxic potential of extremely hydrophobic polychlorinated biphenyl congeners in biota collected at a
Superfund site contaminated with Aroclor 1268. Environ Sci
Technol 32:1214–1221
Khim JS, Villeneuve DL, Kannan K, Lee KT, Snyder SA, Koh CH,
Giesy JP (1999) Alkylphenols, polycyclic aromatic hydrocarbons
(PAHs) and organochlorines in sediment from Lake Shihwa,
Korea: instrumental and bioanalytical characterization. Toxicol
Environ Chem 18: 2424–2432
Meyers-Schöne L, Walton BT (1994) Turtles as monitors of chemical
contaminants in the environment. Rev Environ Contam Toxicol
135:93–153
Nakamura H, Matsuda M, Wakimoto T (1993) Simultaneous determination of several organochlorine compounds (PCDDs/PCDFs,
PCBs, DDTs, HCHs) in limited human samples. J Environ Chem
3:450–451
Olafsson PG, Bryan AM, Stone W (1983) Snapping turtles—a biological screen for PCBs. Chemosphere 12:1525–1532
Podreka S, Georges A, Maher B, Limpus CJ (1998) The environmental
contaminant DDE fails to influence the outcome of sexual
differentiation in the marine turtle Chelonia mydas. Environ
Health Perspect 106:185–188
Ryan JJ, Lau BPY, Hardy JA, Stone WB, O’Keefe P, Gierthy JF (1986)
2,3,7,8-Tetrachlorodibeno-p-dioxin and related dioxins and furans
in snapping turtle (Chelydra serpentina) tissues from the Upper St.
Lawrence River. Chemosphere 15:537–548
Seigel RA, Brauman R (1994) Food habits of the yellow-blotched map
turtle (Graptemys flavimaculata). Report of the US Fish and
Wildlife Service and the Mississippi Department of Wildlife,
Fisheries and Parks
Seigel RA, Brauman R (1995) Reproduction and nesting of the
yellow-blotched map turtle. Report of the US Fish and Wildlife
370
Service and the Mississippi Department of Wildlife, Fisheries and
Parks
Tanabe S (1988) PCB problems in the future: foresight from current
knowledge. Environ Pollut 50:5–28
Tsuda S, Kawano M, Wakimoto T, Tatsukawa R (1993) Application of
charcoal/silica gel column for analysis of polychlorinated dibenzop-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs).
Chemosphere 27:2117–2122
Van den Berg M, Birnbaum L, Bosveld ATC, Brunström B, Cook P,
Feeley M, Giesy JP, Hanberg A, Hasegawa R, Kennedy SW,
Kubiak T, Larsen JC, van Leeuwen FXR, Liem AKD, Nolt C,
K. Kannan et al.
Peterson RE, Poellinger L, Safe S, Schrenk D, Tillitt D, Tysklind
M, Younes M, Wærn F, Zacharewski T (1998) Toxic equivalency
factors (TEFs) for PCBs, PCDDs, PCDFs for humans and wildlife.
Environ Health Perspect 106:775–792
Willingham E, Crews D (1999) Sex reversal effects of environmentally
relevant xenobiotic concentrations on the red-eared slider turtle, a
species with temperature-dependent sex determination. Gen Comp
Endocri 113:429–435
Yawetz A, Benedek-Segal M, Woodin B (1997) Cytochrome P4501A
immunoassay in freshwater turtles and exposure to PCBs and
environmental pollutants. Environ Toxicol Chem 16:1802–1806