Ecotoxicology (2011) 20:1801–1812 DOI 10.1007/s10646-011-0716-z
Metal accumulation and evaluation of effects in a freshwater turtle Shuangying Yu • Richard S. Halbrook • Donald W. Sparling • Robert Colombo
Accepted: 7 June 2011 / Published online: 18 June 2011 Ó Springer Science+Business Media, LLC 2011
Abstract A variety of contaminants have been detected in aquatic and terrestrial environments around the Paducah Gaseous Diffusion Plant (PGDP), Kentucky. The presence of these contaminants at the PGDP may pose a risk to biota, yet little is known about the bioaccumulation of contaminants and associated effects in wildlife, especially in aquatic turtles. The current study was initiated to evaluate: (1) the accumulation of heavy metals (Cd, Cr, Cu, Pb, and Hg) in aquatic ecosystems associated with the PGDP using red-eared slider turtle (Trachemys scripta elegans) as biomonitors; (2) maternal transfer of heavy metals; and (3) potential hematological and immunological effects resulting from metal accumulation. A total of 26 turtles were collected from 7 ponds located south, adjacent, and north of the PGDP. Liver Cu concentrations were significantly different among ponds and Cu concentrations in eggs were positively correlated with female Cu concentrations in kidney. The concentrations of heavy metals measured in turtle tissues and eggs were low and, based on previous studies of reptiles and established avian threshold levels of heavy metals, did not appear to have adverse effects on aquatic turtles inhabiting ponds near the PGDP. However,
S. Yu (&) R. S. Halbrook D. W. Sparling Cooperative Wildlife Research Laboratory, Southern Illinois University Carbondale, Carbondale, IL, USA e-mail:
[email protected] Present Address: S. Yu The Institute of Environmental and Human Health, Texas Tech University, Lubbock, TX, USA R. Colombo Biological Sciences Department, Eastern Illinois University, Charleston, IL, USA
total white blood cell counts, heterophil to lymphocyte ratio, and phytohemagglutinin stimulation index were correlated with metal concentrations. Because other factors may affect the hematological and immunological indices, further investigation is needed to determine if these effects are associated with metal exposure, other contaminants, or disease. Keywords Heavy metals Paducah gaseous diffusion plant Red-eared slider turtle Hematocrit White blood cell count T-cell mediated immunity
Introduction Turtles possess several advantages as biomonitors of environmental contamination compared to many other species. They are widely distributed and occupy a variety of habitats, their long life spans allow monitoring of longterm trends in environmental pollutants, and turtles have sufficient tissue mass for multiple endpoint measurements (Meyers-Scho¨ne et al. 1993). In addition, carnivorous species may accumulate greater concentrations of hazardous chemicals through trophic transfer, and some species are relatively sedentary, making them useful for monitoring contaminants within a specific area. Both aquatic and terrestrial species have been reported as useful biomonitors of metal contamination in the environment. For example, box turtles (Terrapene carolina; Beresford et al. 1981) and common snapping turtles (Chelydra serpentina; Overmann and Krajicek 1995) collected from lead contaminated areas had significantly greater lead concentrations in tissues compared with turtles from reference sites. In addition, there have been numerous studies on contaminant accumulation in sea turtles, which have provided data regarding
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marine pollution (e.g., Franzellitti et al. 2004; Gardner et al. 2006; Maffucci et al. 2005; Storelli et al. 2005). Turtle eggs are also useful biomonitors of environmental contamination because contaminants can be transferred to eggs from female turtles (e.g., Hillestad et al. 1974; Stoneburner et al. 1980) or the ambient environment as demonstrated by studies using lizards (e.g., Brasfield et al. 2004; Marco et al. 2004). Turtle eggs have been used to evaluate contamination of heavy metals (Tryfonas et al. 2006), PAHs (Holliday et al. 2008), PCBs (Dabrowska et al. 2006; Kelly et al. 2008), organochlorine pesticides and dioxins/furans (de Solla and Fernie 2004; Henny et al. 2003). Estimating concentrations of contaminants using eggs has been suggested as a potential non-lethal way to monitor metal exposure for endangered marine turtles (Sakai et al. 1995). Although turtles have been used as biomonitors of environmental contamination, toxicological effects of contaminants are not as well studied in turtles, or other reptilian species, as they have been in fish, birds, or mammals (Sparling et al. 2010). A few studies have reported negative effects in turtles due to exposure to contaminants. Eisenreich et al. (2009) reported that juvenile common snapping turtles exposed to maternally derived PCBs had higher mortality rates than turtles from reference sites at approximately 8 months after hatching. Similarly, Meyers-Scho¨ne (1989) reported significantly higher levels of DNA strand breaks in common snapping turtles and yellow-bellied sliders (Trachemys scripta) with greater concentrations of radionuclides and mercury than turtles from reference sites. With regard to heavy metals, lead influenced survival and righting ability of hatchling slider turtles (T. scripta elegans) injected with 0.25, 1, and 2.5 mg/g lead; survival declined and time to right increased as a function of lead dose (Burger et al. 1998). Similarly, blood mercury concentrations measured in loggerhead sea turtles (Caretta caretta) were negatively correlated with B-lymphocyte proliferation and total white blood cell counts (Day 2003). In order to better understand how contaminants affect reptile species, the current study was initiated to address questions regarding heavy metal accumulation and associated effects in aquatic turtles. The Paducah Gaseous Diffusion Plant (PGDP), a superfund site, was chosen as the study area. It is owned by the U.S. Department of Energy (DOE) and began operation in 1952 to produce enriched uranium for use as commercial nuclear reactor fuel. Plant operations have generated radioactive and other hazardous wastes, including isotopes of uranium, technetium-99, trichloroethylene, PCBs, volatile organic chemicals, and heavy metals (ATSDR 2002). These contaminants have been detected in both terrestrial and aquatic ecosystems on and off the plant. Among the contaminants, heavy metals were detected in outfalls at the PGDP, and the Kentucky Pollutant Discharge Elimination System effluent
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permit limits were occasionally exceeded for heavy metals including Cd, Pb, Cu, and Cr (Phipps and Kzsos 1996). In addition, Hg concentrations in some fish samples collected from Big Bayou Creek downstream of the PGDP were greater than in fish from reference sites (Kszos et al. 1994). Because the presence of these metals at the PGDP may pose a risk to aquatic biota, toxicological studies are needed to assess their bioavailability and potential environmental risk. Red-eared slider turtles, the most abundant aquatic turtle species at the PGDP, were used to evaluate metal accumulation and effects. Metal concentrations were measured in tissues of red-eared slider turtles collected from ponds near the PGDP and their potential effects on hematology and immune function were evaluated. Our specific objectives were to: (1) determine the accumulation of heavy metals (Cd, Cr, Cu, Pb, and Hg) in red-eared slider turtles associated with the PGDP; (2) evaluate maternal transfer of heavy metals; and (3) evaluate the relationship between heavy metals and hematological and immunological indicators, including hematocrit, total and differential blood cell counts, and T-cell mediated immunity. We hypothesized that: (1) based on ground water flow, concentrations of Cd, Cr, Cu, Pb, or Hg in red-eared slider turtles would be greater in ponds adjacent and north of the plant relative to ponds south of the plant; (2) evaluated metals would be detected in eggs due to maternal transfer and the concentrations in eggs would be positively correlated with the concentrations in female tissues; and (3) hematological and immunological indices would be correlated with heavy metal concentrations.
Materials and methods Study sites The PGDP is approximately 15 km west of Paducah, Kentucky, and 5 km south of the Ohio River (Fig. 1). The PGDP facility covers approximately 540 ha; 300 ha within a fenced security area and the remaining 240 ha maintained by the DOE as a buffer zone (USEPA 1995). Approximately 850 ha of land beyond the buffer zone is leased by DOE to the Commonwealth of Kentucky as part of the West Kentucky Wildlife Management Area, and is used for outdoor recreation such as hunting and fishing (USEPA 1995). The PGDP may pose a risk to groundwater due to leaching and dissolution of contaminants from the solid waste management units (Jacobs Engineering Group Inc. 1995). Ponds near the PGDP may be contaminated via groundwater, which flows north to the Ohio River. Seven ponds located on the West Kentucky Wildlife Management Area near the PGDP were sampled during the current study (Fig. 1). Fireplug (FP) and Stick Ponds (ST) were located
Metal accumulation and evaluation of effects in a freshwater turtle
Fig. 1 Map of study ponds near the Paducah Gaseous Diffusion Plant (PGDP), Kentucky. MT Metzger Pond, TP Teal Pond, RL Rock Levee Pond, BP Beaver Pond, DA Disabled Access Pond, ST Stick Pond, FP Fireplug Pond
south, Disabled Access (DA) and Beaver Ponds (BP) adjacent, and Rock Levee (RL), Teal (TP), and Metzger Ponds (MT) were located north of the plant. Three ponds, Metzger, Beaver, and Fireplug, were relatively shallow with soft mud bottoms, whereas Rock Levee, Teal, Disabled Access, and Stick Ponds were relatively deep with hard or rocky bottoms. Fallen tree trunks were present at Beaver, Stick, and Fireplug Ponds, and were most abundant at Beaver Pond, while submerged vegetation was observed in Beaver and Fireplug Ponds. Human recreational activities, such as fishing, were observed more frequently at Metzger, Disabled Access, and Stick Ponds compared to other ponds. Sample collection Baited hoop nets were set at 6 ponds during August 2007 (Teal Pond was not sampled during 2007) and 7 ponds during May through August 2008 to capture red-eared sliders. Captured turtles were weighed (±1 g) using a digital balance, and their straight carapace length and width (±1 mm) as well as plastron length were measured using a digital caliper. Male and female turtles were classified as adults if they had a plastron length greater than 10 cm
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(Tucker et al. 2008). Males were identified by elongation of the foreclaws and longer preanal tail length (Ernst et al.1994; sex was further confirmed by gonads during necropsy). Turtles were transported in coolers to the Cooperative Wildlife Research Laboratory Annex at Southern Illinois University where they were euthanized by decapitation. Total body, liver, and kidney weights (±0.01 g) were recorded, and liver, kidney and approximately 6–16 g of muscle tissue were collected for metal analysis. All tissues were wrapped in aluminum foil, and stored at -20°C prior to analysis. All procedures were approved by the Southern Illinois University Institutional Animal Care and Use Committee (IACUC # A-3078-01). During May through June 2008, captured adult female turtles were X-rayed, and gravid females were kept in wading pools filled with water and sand. Water and food were provided ad libitum. To collect eggs, gravid females were given an intramuscular injection of 1 IU of oxytocin/ 100 g body weight to induce egg laying (Ewert and Legler 1978). Shelled eggs were also collected from the ovaries of female turtles during necropsy. Collected eggs were weighed (±0.01 g), and the greatest length and diameter (±0.1 mm) measured. Eggs were rinsed with tap water followed by deionized water, air dried, and stored in chemically cleaned glass jars at -20°C prior to metal analysis. During 2008, two sediment samples (*100 g) were collected using a core auger along the shoreline of the study ponds following EPA Standard Operating Procedure 2016 (USEPA 1994). Hematology and immune function T-cell mediated immunity was evaluated using the phytohemagglutinin stimulation index (PHA SI; Keller et al. 2006) based on the PHA skin test, which has been widely used to evaluate T-cell mediated immunity in birds (Smits et al. 1999) and reptiles (Belliure et al. 2004; Berger et al. 2005). As a T-cell mitogen, PHA stimulates T-lymphocyte proliferation, differentiation, and cytokine production, resulting in inflammation and swelling of the skin at the injection site (Grasman 2002). Depressed swelling response indicates suppressed T-cell mediated immunity. Fifty microliters of 1 mg/ml PHA (Sigma, St. Louis, Missouri, USA) in phosphate-buffered saline (PBS, pH = 7.2) were injected in the right rear toe web of captured turtles. The left rear toe web served as a control and was injected with 50 ll PBS only. Prior to injection and 24 h post-injection, web thickness was measured using a pressure-sensitive thickness gauge (±0.001 mm, Mitutoyo, Aurora, Illinois). The difference in web thickness of the PHA inoculated web minus the difference in the web thickness of the PBS only inoculated web was used as the PHA SI (Belliure et al. 2004).
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During necropsy, blood was collected from the neck immediately after the turtle was decapitated. To measure hematocrit, blood was drawn into capillary tubes, centrifuged, and the ratio of packed cells to whole blood measured. For WBC counts, blood smears were prepared, air-dried, and fixed in absolute ethanol. Blood smears were stained using Wright-Giemsa stain at the Southern Illinois University School of Medicine, and total and differential WBC counts determined. Total WBC counts were determined as the frequency of WBCs per 1,000 erythrocytes that were randomly observed on each slide (Lajmanovich et al. 2005). Differential WBC counts were determined by counting 200 WBCs and recording proportions of lymphocyte, monocyte, eosinophil, basophil, and heterophil (Casal and Oro´s 2007). The ratio of the number of heterophils to the number of lymphocytes (HL ratio), which has been suggested as an indicator of contaminant exposure in avian species, was calculated (Grasman 2002).
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Total Hg concentrations in sediment samples were analyzed by the Illinois Sustainable Technology Center, Champaign, Illinois. Sediment samples were dried at room temperature (20°C) in a fume hood to a constant mass. A 0.250 g aliquot was digested with 10 ml of double distilled HNO3 and 1 ml of 30% H2O2 in a microwave (Ethos EX microwave extraction labstation, Milestone, Inc., Sheldon, Connecticut, USA) for 10 min at 170°C and 30 min at 180°C. Samples were removed from the microwave and filtered with ashless filter paper to remove any particles in the sample. Samples were then oxidized with 250 ll HCl and 250 ll BrCl, and diluted to 50 ml. Oxidants were neutralized with hydroxylamine immediately prior to analysis. Total mercury concentrations were determined using a PSA Millennium Merlin CVAFS mercury analyzer (P.S. Analytical Inc., Deerfield Beach, Florida, USA). The detection limit was 1 lg/l and the average percent recovery in fortified samples was 96%. All results were reported on a dry weight basis (mg/kg).
Chemical analyses Statistical analyses Homogenized liver, kidney, and muscle tissue, egg contents, and sediment samples were analyzed for Cd, Cr, Cu, and Pb following EPA Method 200.9 (USEPA 2001). All samples were dried at 60°C to a constant mass. Approximately 0.51.0 g of tissue or egg samples was digested with 1.25 ml nitric acid (trace metal grade) at 58°C for 2 h. The resultant solutions were diluted to 25 ml with deionized water and filtered with ashless filter paper to remove lipid in the digested sample before metal analysis. Approximately 1 g of sediment was digested with 10 ml trace metal grade nitric acid at 85°C for at least 2 h, and diluted to 50 ml. Metal concentrations were determined using a Perkin Elmer Zeeman 4100 ZL graphite furnace atomic absorption spectrometer (Perkin–Elmer Corp., Cupertino, California). The method detection limits (MDLs) were 0.94 lg/l (Cd), 1.37 lg/l (Cr), 3.35 lg/l (Cu), and 1.88 lg/l (Pb). The percent recoveries in fortified samples were 77.9% (Cu), 79.4% (Cr), 138% (Pb), and 83.6% (Cd). All results were reported on a dry weight basis (mg/kg). Total mercury concentrations in homogenized liver and egg samples were determined following EPA method 245.7 (USEPA 2005). Approximately 0.1 g of liver tissue or 0.3 g of egg content were digested with 1 ml nitric and 2 ml sulfuric acid (trace metal grade) at 80°C for 30 min, followed by oxidization with potassium permanganate and potassium persulfate at 80°C for 45–60 min. Oxidants were neutralized with hydroxylamine before analysis. Total mercury concentrations were determined using a Hydra AF cold vapor atomic fluorescence spectrometer (Teledyne Leeman Labs, Inc., Hudson, New Hampshire). The MDL for Hg was 0.32 lg/l, and the average percent recovery in fortified tissues was 74.9%.
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For statistical analyses, if C50% of the samples of a tissue had detectable concentrations of a metal, then the concentrations that were below MDL were replaced with concentrations calculated based on the distribution of the data using maximum likelihood estimation methods (Villanueva 2006). Microsoft Excel’s Solver program was used to generate new values for concentrations below MDL (Villanueva 2006). If \50% of the samples of a tissue had detectable concentrations of a metal, then that metal was excluded from statistical analyses for that tissue. Data were analyzed using SPSS 13.0 (SPSS Inc.). Normality and homogeneity of variances were tested using the Kolmogorov–Smirnov statistic and Levene statistic, respectively. Those not conforming to normality were log-transformed and re-tested. Sex differences in metal concentrations were evaluated using a Student’s t test, and a one-way ANOVA or a Welch’s ANOVA was used to determine the differences in concentrations of each metal in each tissue among ponds, followed by Tukey’s HSD, if a significant difference was indicated. Metal concentrations and physiological parameters, including hematological parameters and PHA SI, were examined using Pearson correlation test. Results were considered significant at a = 0.05.
Results Metal accumulation in tissues and sediment Six and twenty red-eared sliders were collected from study ponds during 2007 and 2008, respectively (21 female and 5
Metal accumulation and evaluation of effects in a freshwater turtle
male turtles, 25 adult and 1 juvenile turtles). Age difference was not evaluated because there was only one juvenile turtle. Only one red-eared slider turtle was collected from Rock Levee Pond, thus this pond was excluded from comparisons of metal concentrations among ponds. Chromium and Cu were detected in all tissue samples, and Hg was detected in all liver samples. Lead concentrations were [MDL in 92% liver samples (24/26), 61% kidney samples (14/23), and 52% muscle samples (13/25). Twenty-three percent of liver samples (6/26) had Cd concentrations that were [MDL, while 96% of kidney samples (22/23) had Cd concentrations [ MDL, and no turtles had muscle Cd concentrations [ MDL. Therefore, Cd concentrations in liver and muscle tissues were not evaluated. The greatest mean concentrations of Cr and Pb were measured in liver, whereas mean Cd concentration was greatest in kidney (Table 1). Copper concentrations were similar in liver and kidney tissues, which were greater than concentrations in muscle. The mean kidney Cr concentration
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was significantly greater in male (4.499 ± 1.428 mg/kg) than in female turtles (1.772 ± 0.542 mg/kg; log-transformed, Student’s t = 2.342, df = 21, P = 0.029). No other tissue metal concentrations were significantly different between male and female red-eared slider turtles. Copper concentrations in liver were significantly greater in turtles from Beaver and Metzger Ponds compared with turtles from Teal Pond (F5, 19 = 4.477, P = 0.007; Table 1). Tissue concentrations of Cd, Cr, Pb, and Hg were not significantly different among ponds. The mean Cd and Cr concentrations in sediment ranged from 0.060 to 0.210 mg/kg dw and 2.074 to 3.395 mg/kg among ponds, respectively (Table 2). The mean Cu concentrations in sediment ranged from 1.994 mg/kg (Teal Pond) to 4.802 mg/kg (Beaver Pond). Sediment from Teal Pond had the lowest Pb concentration (3.345 mg/kg) while sediment from Rock Levee Pond had the highest Pb concentration (4.856 mg/kg). The mean Hg concentrations in sediment ranged from 0.029 mg/kg (Beaver Pond) to
Table 1 Metal concentrations (mean ± SE, mg/kg dw) in tissues of red-eared slider turtles (Trachemys scripta elegans) collected during August 2007 and May through August 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky Ponda
Tissue
FP
ST
DA
BP
RL
TP
MT
N
Cd
Liver
3
–b
1.953 ± 0.511
12.815 ± 5.245AB
1.123 ± 0.124
0.532 ± 0.187
Kidney
2
1.031 ± 0.183
0.666 ± 0.191
14.996 ± 3.996
\MDL
NMc
Muscle
2
–
1.386 ± 0.128
5.314 ± 1.862
\MDL
NM
Liver
3
–
4.128 ± 1.927
9.165 ± 4.253AB
0.935 ± 0.799
0.708 ± 0.357
Cu
Pb
Hg
Kidney
3
0.571 ± 0.122
2.550 ± 1.031
12.628 ± 1.240
0.352 ± 0.200
NM
Muscle
3
–
1.146 ± 0.474
2.690 ± 0.578
0.813 ± 0.395
NM
Liver
5
–
3.046 ± 0.915
7.602 ± 1.495AB
1.524 ± 1.204
0.555 ± 0.166
Kidney
3
0.331 ± 0.150
4.812 ± 2.016
16.943 ± 8.062
\MDL
NM
Muscle
5
–
1.316 ± 0.497
6.907 ± 4.160
0.022 ± 0.003
NM
Liver
3
–
7.392 ± 2.860
22.606 ± 2.320A
1.474 ± 0.391
1.198 ± 0.323
Kidney Muscle
3 3
0.993 ± 0.318 –
0.843 ± 0.059 1.035 ± 0.444
14.876 ± 2.544 3.431 ± 0.438
0.897 ± 0.612 0.274 ± 0.126
NM NM
Liver
1
–
1.410
7.202
\MDL
0.255
Kidney
1
0.238
7.465
10.32
0.921
NM
Muscle
1
–
2.481
1.748
0.729
NM 0.489 ± 0.152
Liver
6
–
2.430 ± 1.174
6.018 ± 0.836B
0.470 ± 0.081
Kidney
6
0.523 ± 0.115
0.787 ± 0.194
10.130 ± 1.417
0.142 ± 0.042
NM
Muscle
6
–
2.096 ± 0.556
3.945 ± 1.292
0.283 ± 0.102
NM
Liver
5
–
2.943 ± 0.506
20.492 ± 6.103A
2.025 ± 0.980
1.083 ± 0.275
Kidney
5
0.814 ± 0.225
2.708 ± 1.506
19.111 ± 3.778
0.605 ± 0.359
NM
5
–
1.430 ± 0.328
3.866 ± 1.012
0.380 ± 0.213
NM
26
–
3.321 ± 0.559
12.213 ± 1.829
1.111 ± 0.255
0.719 ± 0.100
Muscle Total
Cr
Liver Kidney
23
0.661 ± 0.085
2.247 ± 0.543
14.347 ± 1.463
0.378 ± 0.121
NM
Muscle
25
–
1.524 ± 0.197
4.331 ± 0.895
0.327 ± 0.078
NM
Concentration means followed by different uppercase letters are significantly different (Student’s t test, P \ 0.05) a
MT Metzger Pond, TP Teal Pond, RL Rock Levee Pond, BP Beaver Pond, DA Disabled Access Pond, ST Stick Pond, FP Fireplug Pond
b
\50% of the samples of a tissue had detectable Cd concentrations
c
Not measured
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Table 2 Mean concentrations of heavy metals (mg/kg dw, n = 2) in sediment collected during July 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky Ponda
Cd
Cr
Cu
Pb
Hg
FP
0.065 (0.039, 0.091)b
3.395 (2.953, 3.837)
4.738 (4.604, 4.873)
3.981 (3.790, 4.172)
0.053 (0.046, 0.060)
ST
0.210 (0.204, 0.215)
2.968 (2.947, 2.988)
4.240 (3.735, 4.745)
4.339 (3.746, 4.931)
0.035 (0.034, 0.036)
DA
0.060 (0.060, 0.060)
2.887 (2.741, 3.033)
4.380 (4.276, 4.485)
4.630 (3.928, 5.332)
0.033 (0.024, 0.043)
BP
0.069 (0.064, 0.074)
2.522 (2.035, 3.009)
4.802 (4.341, 5.262)
3.711 (3.674, 3.749)
0.029 (0.027, 0.031)
RL
0.103 (0.100, 0.105)
2.878 (2.457, 3.299)
4.506 (4.463, 4.549)
4.856 (4.062, 5.649)
0.054 (0.053, 0.055)
TP
0.081 (0.066, 0.095)
2.393 (2.203, 2.583)
1.994 (1.692, 2.296)
3.345 (2.296, 4.394)
NMc
MT
0.203 (0.162, 0.244)
2.074 (1.940, 2.207)
4.598 (4.070, 5.125)
4.820 (4.266, 5.374)
0.054 (0.048, 0.059)
a
FP fireplug pond, ST stick pond, DA disabled access pond, BP beaver pond, RL rock levee pond, TP teal pond, MT metzger pond
b
Concentrations of metals in each sediment sample
c
Not measured
0.054 mg/kg (Metzger and Rock Levee Ponds). Mercury concentration in sediment was not measured for Teal Pond. Maternal transfer A total of 56 eggs were collected from 10 female red-eared sliders for metal analysis (2–12 eggs per female). Fifty-one eggs were analyzed for Cd, Cr, Cu, and Pb, and 40 eggs for Hg. Of the 5 metals examined, Hg was [MDL in 32 (80%) of the eggs (mean = 0.054 ± 0.006 mg/kg dw) and Cu was [MDL in 51 (100%) of the eggs (mean = 2.650 ± 0.118 mg/kg dw). Concentrations of Cr and Cd were \MDL in all eggs, and Pb concentration was \MDL in 90% of the eggs. Therefore, maternal transfer of Cd, Cr, and Pb was not evaluated. For Cu and Hg, the average concentration of eggs from each clutch was calculated and used to evaluate correlations between concentrations in eggs and females. Copper concentrations in eggs were positively correlated with female Cu concentrations in kidney (Pearson’s r = 0.748, P = 0.033; Fig. 2), but not with Cu concentrations in liver (r = 0.550, P = 0.125) or muscle (r = 0.308, P = 0.458). Mercury concentrations in eggs were not correlated with female liver Hg concentrations (r = 0.377, P = 0.284). Hematology and immune function Lymphocytes were the most commonly observed WBC in red-eared slider blood, followed by eosinophils, heterophils, and monocytes, while basophils were relatively rare compared with other white blood cell types. The relationships between metal concentrations and hematocrit and WBC counts were evaluated in 21 and 17 of the 26 turtles evaluated for metal accumulation, respectively. Total WBC counts were negatively correlated with liver Pb concentrations (Pearson’s r = -0.495, P = 0.043, Fig. 3). However, a female turtle from Metzger’s Pond had outlying
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Fig. 2 Correlation of Cu concentrations measured in kidney tissue of female red-eared slider turtles (Trachemys scripta elegans) collected during May through August 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky and in their eggs (P = 0.033, n = 8)
value (5.744 mg/kg) indicated by Grubb’s test (Z = 2.62, P \ 0.05). After this outlier was removed, TWBC was still correlated with liver Pb concentrations (Pearson’s r = -0.601, P = 0.014). Total WBC counts were positively correlated with log-transformed kidney Cr concentrations (r = 0.707, P = 0.002, Fig. 4). Log-transformed HL ratios decreased significantly with increased kidney Cd concentrations (r = -0.707, P = 0.007, Fig. 5), however, one turtle from Teal Pond had an outlying value (1.63; Grubb’s test, Z = 2.62, P \ 0.05). The outlier was removed from the correlation analysis and the data re-tested. Upon re-evaluation, the HL ratios were still negatively correlated with kidney concentrations (r = -0.693, P = 0.013). Hematocrit was not correlated with any metal concentrations in any tissues.
Metal accumulation and evaluation of effects in a freshwater turtle
Fig. 3 Correlation of Pb concentrations measured in liver tissue of red-eared slider turtles (Trachemys scripta elegans) collected during August 2007 and May through August 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky and total white blood cell (WBC) counts of these turtles (P = 0.043, n = 17)
Fig. 4 Correlation of log-transformed Cr concentrations measured in kidney tissue of red-eared slider turtles (Trachemys scripta elegans) collected during August 2007 and May through August 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky and total white blood cell (WBC) counts of these turtles (P = 0.002, n = 17)
Phytohemagglutinin SI was measured in 18 red-eared sliders. The SIs were negative for 3 of these turtles, and therefore these values were excluded from statistical analysis. Upon initial evaluation, the SIs were positively correlated with Cu concentrations in kidney (r = 0.628, P = 0.012; Fig. 6); however, Grubb’s test indicated an outlying value (Z = 2.5073, P \ 0.05) in one turtle from Disabled Access Pond with a high PHA SI value (0.422). After the outlier was excluded, the correlation was not
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Fig. 5 Correlation of Cd concentrations measured in kidney tissue of red-eared slider turtles (Trachemys scripta elegans) collected during August 2007 and May through August 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky and log-transformed ratios of heterophils to lymphocytes of these turtles (P = 0.007, n = 17)
Fig. 6 Correlation of Cu concentrations measured in kidney tissue of red-eared slider turtles (Trachemys scripta elegans) collected during August 2007 and May through August 2008 from ponds near the Paducah Gaseous Diffusion Plant, Kentucky, and the phytohemagglutinin (PHA) stimulation index of these turtles (P = 0.012, n = 15)
significant (r = 0.297, P = 0.303). There was no correlation between other metals in other tissues and PHA SI.
Discussion Metal accumulation in tissues and sediment The mean liver concentrations of Cd (\0.118 mg/kg dw), Cr (3.321 ± 0.559 mg/kg), or Cu (12.213 ± 1.829 mg/kg)
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in red-eared sliders collected near the PGDP were lower or comparable to those in turtles from reference sites (Cd 0.17 ± 0.08 mg/kg dw, Cr 1.16 ± 0.31 mg/kg, and Cu 52.73 ± 2.12 mg/kg) reported by Nagle et al. (2001). Similarly, metal concentrations measured in the current study were lower or comparable to those reported in common snapping turtles (Albers et al. 1986; Overmann and Krajicek 1995). Because little information is available on threshold concentrations of metals in tissues for reptiles, metal concentrations measured in the current study were compared with concentrations associated with adverse effect in avian species. Copper and Cr were not compared because they are essential elements and few studies have reported their threshold concentrations. For Cd, red-eared sliders collected near the PGDP had low concentrations (maximum tissue concentration 0.296 mg/kg ww) compared with 40 mg/kg in the liver or 100 mg/kg in the kidney suggested by Furness (1996) as the threshold levels in birds. Similarly, the greatest concentration of Pb measured in the current study (1.315 mg/kg ww in liver tissue) was also lower than liver Pb concentrations [2 mg/kg ww which may cause subclinical poisoning in waterfowl (Pain 1996). The greatest Hg concentration measured in turtles in the current study (0.513 mg/kg ww) was considerably lower than that suggested by Zillioux et al. (1993; 5 mg/kg ww in liver) as the conservative threshold for detrimental effects in waterfowls. Our results did not support the hypothesis that turtles collected from ponds adjacent and north of the plant have greater metal concentrations than turtles from ponds south of the plant. Significant difference in metal concentrations among ponds was only detected for Cu. Copper concentrations measured in liver tissues of turtles collected from Teal Pond were significantly lower than concentrations measured in liver tissues of turtles from Metzger and Beaver Ponds. Although not significantly different, turtles from Teal Pond also had the lowest mean Cu concentration in kidney (Table 1), which is another primary organ of metal accumulation. This difference of Cu accumulation may be due to the age of Teal Pond rather than an effect associated with location relative to the PGDP. Metzger Pond is also located north of the PGDP, near Teal Pond, and Cu concentrations in liver tissues of turtles collected from Metzger Pond were similar to concentrations measured in liver tissues of turtles from Beaver Pond which is located adjacent to the PGDP. Teal Pond was built in May 2007 and is relatively new compared to other ponds that have been in the study area for approximately 10–20 years. Analysis of sediment metal concentrations also indicated that sediment from Teal Pond contained lower Cu concentrations (1.994 mg/kg dw) compared to sediment from other ponds (ranged from 4.240 to 4.802 mg/kg).
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Concentrations of other heavy metals were not significantly different among ponds. It is possible that turtle immigration and emigration may affect the comparison of metal concentrations among ponds in the current study. However, during a concurrent study of turtle population in the study area (Yu 2009), among the 162 red-eared slider turtles captured in 2007, only 5 of them (3%) were captured in 2008 at different ponds from their original pond of capture. Therefore, turtle movement does not appear to significantly influence comparison of contaminant data among ponds. In the current study, kidney Cr concentration was the only tissue metal concentration that was significantly different between male and female turtles. Besides the difference in physiological requirement of Cr associated with sex, it is also possible that females eliminated Cr via eggs and thus had lower concentrations (Meyers-Scho¨ne and Walton 1994). However, Cr concentrations were below MDL in eggs in the current study. Previous studies have reported inconsistent results regarding sex-related difference in metal accumulation. Male and female loggerhead turtles did not differ in concentrations of Cd, Cu, Fe, Mn, Ni, and Zn (Franzellitti et al. 2004), whereas Blanvillain et al. (2007) reported that female diamondback terrapins (Malaclemys terrapin) had greater Hg concentrations than males. Maternal transfer of heavy metals In reptiles, females deposit trace elements in eggs (Linder and Grillitsch 2000; Smith et al. 2007). Studies have demonstrated that both essential (e.g., Zn, Se, Cu, and Cr) and non-essential elements (e.g., Cd, Pb, and Hg) are deposited in reptile eggs through maternal transfer (Burger 2002; Guirlet et al. 2008; Hopkins et al. 2004; Roe et al. 2004; Tryfonas et al. 2006). In the current study, while concentrations of Cr, Cd, and Pb were low in most of the eggs, Cu was measured in all the eggs evaluated. The mean Cu concentration in eggs in the current study (2.65 ± 0.12 mg/kg dw) was greater than concentrations reported in red-eared slider eggs from the lower Illinois River (0.9 ± 0.5 mg/kg dw; Tryfonas et al. 2006). The positive relationship between Cu concentrations in female kidney tissue and eggs detected in the current study suggests that egg Cu concentration may be useful as an indicator of female Cu exposure. Although Cu concentrations in eggs were not correlated with Cu concentrations in female liver tissue, the power of the correlation (0.34) is low and it is likely that small sample size may result in the insignificant correlation. Future studies with large sample size are needed to determine the relationship between metal concentrations in eggs and female tissues. The mean concentration of Hg (0.054 ± 0.006 mg/kg dw) measured in eggs collected during the current study is
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comparable to those reported in eggs of slider turtles from the Savannah River Site, South Carolina (0.04 ± 0.015 mg/kg dw; Burger and Gibbons 1998). In the current study, egg Hg concentrations were not correlated with female liver Hg concentrations. Similarly, Guirlet et al. (2008) reported no correlation between egg and female blood Hg concentrations in leatherback turtles (Dermochelys coriacea). In the current study, the greatest Hg concentration measured in eggs was 0.156 mg/kg ww, which is lower than the concentrations previously reported to adversely affect early development in avian or reptilian species. Zillioux et al. (1993) reviewed avian laboratory and field data from previous studies and suggested that egg Hg concentrations of about 1 mg/kg ww may cause behavioral effect and concentrations of 5–6 mg/kg may result in mortality and brain lesions. Heinz et al. (2009) evaluated the median lethal concentration (LC50) of MeHg embryotoxicity in 23 avian species, and reported the lowest LC50s were \0.25 mg/kg ww MeHg in highly sensitive species. However, they also noted that injected Hg has a higher toxicity than Hg transferred maternally. Regardless, the lowest LC50 reported by Heinz et al. (2009) that was associated with embryotoxicity was greater than the maximum concentration measured in turtle eggs in the current study. Bishop et al. (1998) reported a range of Hg concentrations from \0.05 to 0.14 mg/kg ww in common snapping turtle eggs, and indicated that Hg concentrations were not correlated with the incidence of embryonic abnormalities. Hematology and immune function Reptiles are declining globally, and environmental contamination has been suggested as one of the major causes (Gibbons et al. 2000); however, toxicological studies involving reptiles are relatively scarce compared to other wildlife (Hopkins 2000; Sparling et al. 2010). Furthermore, most studies report tissue concentrations without reference to potential physiological effects. Without known concentrations associated with toxic effects, it is difficult to interpret analytical data (Storelli and Marcotrigiano 2002). Hematological parameters can be used to evaluate general animal health, and evaluations of the immune system can provide evidence of impaired immune function, both of which were included in the current study. Hematocrit is an index of condition, nutrition, or general health; low hematocrit can be an indication of anemia (Peterson 2002), and some metals, such as Pb, can cause anemia and low hematocrit (Hoffman et al. 1981). In the current study, hematocrit values of red-eared sliders were not correlated with metal concentrations in tissues, suggesting that the measured metal concentrations may not adversely affect hematocrit.
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The negative correlation between total WBC counts and liver Pb concentrations indicates that Pb concentrations measured in the current study may have an immunosuppressive effect in red-eared slider turtles. Decreases in total WBC counts associated with exposure to Pb have also been observed in male mallards (Anas platyrhynchos) exposed to Pb (Rocke and Samuel 1991). However, Pb concentrations measured in the current study were low, and with the small sample size, further investigation is needed to reveal the relationship between immunity and Pb concentrations in aquatic turtles. In contrast, a positive correlation between kidney Cr concentrations and total WBC counts was detected. Chromium is an essential element, however, little is known about potential Cr functions in reptile immune systems. The ratio of heterophils to lymphocytes has been suggested to be a reliable indicator of stress in birds (Gross and Siegel 1983). Increased stress hormone levels influence this ratio by increasing heterophils and decreasing lymphocytes (Gross and Siegel 1983). Increased HL ratios have been observed in avian species exposed to organic chemicals and trace elements including selenium, arsenic, boron, and lead (Grasman 2002) and MeHg (Spalding et al. 2000), and in sea turtles exposed to TCDD-like PCBs (Keller et al. 2004). Although the impact of heavy metals on HL ratio remains unclear in turtles, it is reasonable to suggest that Cd exposure may increase HL ratios because Cd is a highly toxic element. It is possible that other factors may have decreased HL ratios in the current study. First, only 5 heavy metals were examined in the current study, however, turtles collected near the PGDP may be exposed to other contaminants because the PGDP facility has released a variety of contaminants (ATSDR 2002). Keller et al. (2004) reported that the HL ratios ranged from 0.1 to 1.4 in loggerhead turtles, which were positively correlated with concentrations of organochlorine contaminants. The HL ratios measured in the current study ranged from 0.02 to 1.63, similar to those reported by Keller et al. (2004). It is possible that turtles in the current study with low Cd concentrations may have been exposed to organic contaminants that potentially resulted in high HL ratios. Second, although turtles were examined for physical abnormalities when captured, evaluations for disease were not included in this study. Grasman (2002) suggested that one challenge in the interpretation of HL ratio was distinguishing stress response from disease or infections which can also change the WBC profile. In reptiles, heterophil counts increase dramatically in the case of extracellular bacterial infections (Frye 1991). Third, transportation and handling of the turtles may cause stress. Turtles were captured using hoop nets and were measured for physical parameters such as weight and size. Captured turtles were then transported to Southern Illinois University, which took
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approximately an hour. Turtles were kept in captivity for about 5–10 days during which females were X-rayed and egg laying was induced, and all the turtles were examined for T-cell mediated immunity. These procedures may be stressful to the turtles, which could potentially affect the HL ratio. In the current study, one turtle had an outlying PHA SI which when included in statistical evaluations resulted in a positive correlation with Cu concentrations in kidney tissue. However, when this value was removed from statistical analysis, the correlation was no longer significant. The outlier was not believed to be an experimental error and may simply have been the result of individual difference, which would suggest that Cu may be associated with greater immune response in turtles. As an essential element, Cu is a component of some enzymes (Eisler 1998) and may play an important role in immune function (Wintergerst et al. 2007). Copper deficiency may result in impairment of cell-mediated immunity, e.g., reduction in T-lymphocyte numbers in rats and decreased responsiveness of splenic lymphocytes to T-cell mitogens (Bala et al. 1990; Bonham et al. 2002). However, little is known about the effect of Cu on T-cell mediated immunity in wildlife, and further studies are recommended to determine the role of Cu in the immune function of turtles and other reptiles.
Conclusions Concentrations of Cd, Cr, Cu, Pb, and Hg in turtle tissues and eggs measured in the current study were generally low and do not appear to have adverse effects on turtles inhabiting ponds near the PGDP. Statistical analysis did not indicate a significant south-north gradient of metal concentrations in turtles collected near the PGDP. Therefore, we suggest that plant operations may not influence the distribution of heavy metals examined in the current study. The positive correlation of Cu concentrations in female turtles and eggs measured in the current study indicates that Cu can be deposited to eggs via maternal transfer, and egg Cu concentrations could be used as an indicator of Cu concentrations in female turtles. Correlations between hematological and immunological indices and some metal concentrations were detected, indicating that hematological and immunological effects may occur during chronic exposure to heavy metals. However, metal concentrations measured in the current study are low and our sample size is small. In addition, other factors may affect the hematological and immunological indices. Further studies are recommended to determine if these effects are associated with metal exposure, other contaminants, or disease.
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S. Yu et al. Acknowledgments This project was funded by the Kentucky Department of Environmental Protection (MOA 0600004073). We wish to thank the staff at the West Kentucky Wildlife Management Area and the Paducah Gaseous Diffusion Plant for their assistance. We would like to thank Scott Weir, Tricia Trimble, Dawn Fallacara, Laura Pratt, and Maureen Doran for their help with laboratory and field work. Philip Villanueva from the EPA Office of Pesticide Programs kindly provided the Solver program. Comments from two anonymous reviewers greatly improved the manuscript.
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