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Effects of inorganic lead on Western fence lizards (Sceloporus occidentalis)
Environmental Pollution 157 (2009) 3457–3464
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Effects of inorganic lead on Western fence lizards (Sceloporus occidentalis) Christopher J. Salice a, *, Jamie G. Suski a,1, Matthew A. Bazar a, Larry G. Talent b a b
US Army Center for Health Promotion and Preventive Medicine, Aberdeen Proving Ground, MD 21010, USA Oklahoma State University, Department of Natural Resource Ecology and Management, Stillwater, OK 74078, USA
The Western fence lizard, Sceloporus occidentalis, is sensitive to Pb and is a useful laboratory model for ecotoxicological testing of reptiles.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 30 January 2009 Received in revised form 3 June 2009 Accepted 15 June 2009
Although anthropogenic pollutants are thought to threaten reptilian species, there are few toxicity studies on reptiles. We evaluated the toxicity of Pb as lead acetate to the Western fence lizard (Sceloporus occidentalis). The acute lethal dose and sub-acute (14-day) toxicity studies were used to narrow exposure concentrations for a sub-chronic (60-day) study. In the sub-chronic study, adult and juvenile male lizards were dosed via gavage with 0, 1, 10 and 20 mg Pb/kg-bw/day. Mortality was limited and occurred only at the highest dose (20 mg Pb/kg-bw/d). There were statistically significant sub-lethal effects of 10 and 20 mg Pb/kg-bw/d on body weight, cricket consumption, organ weight, hematological parameters and post-dose behaviors. Of these, Pb-induced changes in body weight are most useful for ecological risk assessment because it is linked to fitness in wild lizard populations. The Western fence lizard is a useful model for reptilian toxicity studies. Ó 2009 Elsevier Ltd. All rights reserved.
Keywords: Reptiles Lizard Lead Toxicity Metals
1. Introduction An important objective in ecotoxicology is to generate data that can be used to facilitate environmental management, principally through the ecological risk assessment (ERA) process. Frequently, however, risk assessments must be conducted with incomplete toxicological datasets on receptors of interest and, hence, rely on toxicity data for surrogate species which, can introduce significant uncertainty. For example, U.S. EPA guidance on risk assessment of pesticides suggests that birds can be used as surrogates for reptiles and that fish can be used as surrogate taxa for amphibians (US EPA, 2004). Although risk estimates based on surrogate species are convenient, they may lead to significant errors in projecting risk; risk estimates may be too high or too low resulting in cost-prohibitive overprotection and environmentally damaging under-protection, respectively. At a minimum it would be beneficial to have laboratory models representing each vertebrate class. Ideally, the availability of more than one representative for each class would provide a better perspective on the sensitivity of a given class to a toxicant. While
* Corresponding author. Present address: Texas Tech University, The Institute of Environmental and Human Health, Box 41163, USA. Tel.: þ1 806 885 4567x229; fax: þ1 806 885 2132. E-mail addresses: [email protected] (C.J. Salice), [email protected] (J.G. Suski), [email protected] (M.A. Bazar), [email protected] (L.G. Talent). 1 Present address: Texas Tech University, Department of Biology. 0269-7491/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2009.06.013
there are standard toxicity test species for most animal classes, to date there have been no reptilian species routinely used for toxicity testing. The animal class, Reptilia, has been historically under-represented by the field of toxicology (Hopkins, 2000; Campbell and Campbell, 2002). Despite the fact that reptiles are widespread, there has been relatively few laboratory toxicity studies conducted. Reptiles may have been excluded from consideration because it was believed that benchmarks protective of mammalian and avian receptors were likely also protective of reptilian receptors. However, the approach to using non-reptile toxicity data was necessitated, in part, by the lack of a viable laboratory reptile model, which was likely driven by the fact that many reptiles can be slow to reach reproductive maturity and may not be conducive to convenient laboratory rearing due to size or special husbandry considerations. Also, it is possible that while methods and concepts related to ecotoxicology and ERAs were developing, focus was placed on more prominent and readily available mammalian and avian models. There is, however, a growing interest in assessing risks to historically under-represented species such as those belonging to the class reptilia (Hopkins, 2000; Campbell and Campbell, 2002). In addition, there is evidence that lizard populations are declining and of the six main threats discussed by Gibbons et al. (2000), anthropogenic pollutants are the greatest. The development of a routine reptilian toxicity-testing model could foster the development of reptilian toxicity benchmarks for use in ERAs and ultimately contribute to improved management decisions with regard to naturally occurring reptile populations.
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Previous efforts to develop a reptile toxicity model resulted in the identification of several populations of Eastern (Sceloporus undulatus) and Western (Sceloporus occidentalis) fence lizards that can be reared and bred successfully in the laboratory (Talent et al., 2002). Subsequently, Western fence lizards have been used to evaluate the toxicity of several toxicants including military compounds (Suski et al., 2008; McFarland et al., 2008), pesticides (Talent, 2005; Holem et al., 2006) and endocrine disrupters (Brasfield et al., 2002). The objective of the present study was to build upon previous efforts to further utilize a lizard model for routine toxicity testing. The approach was modeled after standardized mammalian toxicity tests that progress in a hierarchical fashion from acute lethal dose, short-term sub-acute exposures, and longer sub-chronic exposures (Suski et al., 2008). Lead (Pb) is a naturally occurring metal but is not considered an essential nutrient for plants or animals (US EPA, 2005a). It is widely recognized as an important environmental contaminant from industrial activities such as smelting and mining (Zhang et al., 2007; Panichayapichet et al., 2007), and on Small Arms Firing Ranges (SAFRs) from the use of lead shot (ITRC, 2003; US EPA, 2005b). Since it is likely that reptiles inhabiting these areas would be exposed to Pb, reptilian toxicity data could be used in screeninglevel or refined ERAs. Hence, the main objectives of the present study were to evaluate the acute and sub-chronic toxicity of lead to S. occidentalis and to further vet the testing system for potential development as a routine reptilian toxicity bioassay. Acute (single dose), sub-acute (daily dose for 14 days) and sub-chronic (daily dose for 60 days) studies were conducted. In addition the subchronic study was conducted for both juvenile and adult male Western fence lizards to obtain some insight into the potential for age-specific sensitivity that may occur over longer-term exposures. 2. Materials and methods 2.1. Lizards and husbandry Adult (male and female) and juvenile (male) S. occidentalis were obtained from laboratory-reared stocks maintained at Oklahoma State University that were descendents of wild-caught animals from the San Joaquin Valley (CA, USA). The approximate age of lizards was 4 months for juvenile and one year for adults. All lizards were maintained in the laboratory for two weeks prior to study initiation. Pre-study conditions were the same as conditions during the study. General husbandry information for S. occidentalis has been previously published (Talent et al., 2002; Suski et al., 2008; McFarland et al., 2008). Briefly, lizards were maintained individually in modified, clear-plastic mouse cages. The cage dimensions were 20 cm deep, 20 cm wide and 40.5 cm long. Wood chip bedding, approximately 1.3 cm deep, a 10 cm PVC (3.8 cm diameter) and a water dish (small plastic petri dish) were placed in every cage. Water was changed daily, woodchips and cages were changed once per week. Lizards were fed 1–4 house crickets (Acheta domesticus) per day depending on the size of the cricket; lizards on study were given 2, 1.3 cm crickets per day. Crickets were dusted with Rep-CalÒ (Rep-Cal Research Labs, Los Gatos, CA, USA), a vitamin and mineral supplement to help prevent metabolic bone disease. Importantly, many reptiles require a temperature gradient for optimal health. A temperature gradient from 22 to 34 C was maintained in each cage by keeping room temperature at 22 C and using an electric heat strip to warm one end of the cage. Light was ambient fluorescent supplemented with incandescent both on a 12:12 light:dark cycle. Humidity ranged between 30 and 70%. This study was conducted under a Good Laboratory Practices (GLP) protocol approved by United States Army Center for Health Promotion and Preventive Medicine (USACHPPM) Animal Care and Use Committee (IACUC). As with any toxicity study, the quality of results is strongly dependent on accurate dosing technique. The dosing method employed allowed oral dosing of adult lizards using a 1 cc syringe and a 22 gauge stainless steel feeding tube; a pipette may also be used (Suski et al., 2008). Doses were delivered orally because lizards are likely to be exposed to many toxicants via the diet. In addition, oral delivery was a more precise dosing regimen compared to feeding studies. Dosing was found to be more efficient with two people. Standard dosing procedure involved one person obtaining the lizard and while holding it in one hand, gently applying steady downward pressure on the dewlap until the mouth opens. The person administering the dose gently places one end of a small wooden dowel in the corner of the mouth to ensure the mouth does not close. The dose is administered to the rear of the mouth but not past the throat. Unlike in many rodent studies, the dosing
tube was not placed directly into the stomach as this was thought to be too traumatic for lizards. While a pilot study indicated that adult lizards could be dosed with as much as 0.15 ml, we found that dosing success was nearly 100% when dosing volume was between 0.05 and 0.10 ml. 2.2. Test article Lead acetate was obtained from Aldrich chemical company. Because Lead acetate is soluble, distilled water was used as the carrier for dosing experiments; analytical results showed that Pb concentrations in control water would result in doses to lizards of about 0.008 mg/kg, well below the lowest dose tested of 1.0 mg/ kg. All dosing solutions were analyzed for total lead concentration following EPA method 6010B (ICP) using a Perkin Elmer Spectrometer (Optima 3000 XL). Test concentrations were verified prior to and after the experiment. 2.3. Experimental design The approach to developing the Western fence lizard as a viable toxicity-testing model was based largely on standardized approaches for mammalian testing (Suski et al., 2008; McFarland et al., 2008). In the present study, a hierarchy of studies was used to proceed from acute to sub-chronic studies. This methodical approach helped to focus, for a given study duration, the range of toxicant concentrations needed to generate dose-dependent variation in response. In addition, the method ultimately reduced the total number of animals required. 2.4. Acute toxicity study The first step in evaluating the toxicity of lead acetate to lizards was to determine an Approximate Lethal Dose (ALD) of Pb, which is the lowest dose that causes mortality within 14 days following a single administered dose. The method requires using only one animal of each sex at each dose level. Although this method does not result in a statistical estimate of a lethal dose such as an LD50, it approximates the lethal dose while reducing the total number of animals used; the ALD establishes the upper bound for longer duration toxicity studies. Although the primary endpoint of interest in the ALD was mortality, other endpoints were evaluated such as growth, food consumption, behavior and some hematological parameters. Average mass of lizards used for the ALD was 11.3 þ/ 1.14 g for males and 11.9 þ/ 0.87 g for females. All lizards were adults. One lizard of each sex was given a deionized water control, or as lead acetate solutions with one of the following doses of Pb: 100, 200, 500, 1000, 2000, 3000, 4000 and 8000 mg Pb/kg body mass, denoted as mg/kg hereafter. To achieve a dose of 8000 mg Pb/kg body mass, two separate dosing events, separated by 1 h, were required. Pilot studies indicated that this was sufficient time to allow 2 doses near 0.15 ml to be administered and it was unlikely there would be any time-delay effects on toxicity. 2.5. 14-Day sub-acute toxicity study The 14-day study was designed as a range-finding study to further define suitable doses of Pb for use in the longer-term sub-chronic study. Six adult lizards per sex were used for each treatment level and given one dose per day for 14 consecutive days. The doses were 0 (water control), 31.25, 62.50, 125, 250, 500, and 1000 mg Pb/kg body mass/d, denoted as mg/kg/d hereafter. Six adult animals of each sex per dose provided enough insight into potential levels of toxicity while also minimizing the number of animals used. Lizards were observed daily for signs of toxicity including morbidity, ataxia or hyperactivity and apparent dosing-related behaviors (explained below). Each lizard was given 2–4 fresh crickets each day and the number of crickets remaining on the following day was recorded as a measure of food consumption. Lizards were weighed before and during the 14-day study on day 3, 1, 0, 7 and 14. Lizards were sorted into treatment groups using LabcatÒ such that there were no significant differences in weight among groups. Treatments were then randomly assigned to each group. All lizards alive on day 14 were euthanized and necropsied. Whole blood was collected with heparinized capillary tubes for hematology and clinical chemistry analyses. 2.6. 60-Day sub-chronic toxicity study The length of the sub-chronic exposure was 60 days, which is approximately 3–16% of the life span of S. occidentalis (Ruth, 1977). Forty juvenile and forty adult male lizards were randomly sorted into four dose groups. Female lizards were excluded from the sub-chronic toxicity study because they can spontaneously enter vitellogenesis and internal egg masses can account for 20% of body mass, potentially confounding dosing. Based on the results of the 14-day study, doses were selected as 0, 1, 10, or 20 mg/ kg/d. Doses were administered to each lizard, every day for 60 consecutive days. All animals were weighed on day 3, 1, 0, 3, 7 and weekly thereafter. Lizards were fed 2 crickets per day and were allowed 24 h to consume crickets. All uneaten crickets were removed the following day and counted to monitor food consumption. Daily observations were made on condition; if animals appeared lethargic or moribund or did not respond to stimulus or righting, they were euthanized. Any animals
C.J. Salice et al. / Environmental Pollution 157 (2009) 3457–3464 remaining alive on day 60 were euthanized using CO2 followed by decapitation. Whole blood was collected with heparinized capillary tubes for hematological and clinical chemistry analyses. Additionally, brain, liver, heart, kidneys and gonads were removed, weighed and preserved in formalin. Post-dose observations included turning dark in coloration, assuming arched posture and/or mouth gaping were recorded during the 60-d study. Dark coloration could be described as an overall darkening of the skin (from gray to charcoal gray/ black) accompanied by sharply contrasting lighter markings that are less pronounced under normal conditions. Arched posture describes lizards that appear to extend the forelimbs so that the head is facing nearly vertical with the hind limbs and tail resting on the ground. This behavior is different than head-bobbing or push-ups in that there was no cyclical movement; once forelimbs were extended, lizards remained in that position. Gaping was characterized by either, keeping the mouth open or repeated opening and closing. All lizards were observed for occurrence of dark coloration, arched posture or gaping within 30 min after a dose had been administered. 2.7. Hematology and clinical chemistries Following euthanasia, blood samples were collected with several heparinized capillary tubes (lithium heparin). The heparinized sample was then discharged from each capillary tube into a plain 1.5 ml microtube for analysis. Approximately 35 ml of the whole blood sample was drawn off with a plain capillary tube and centrifuged for manual determination of hematocrit and total solids with a card reader and refractometer. 50 ml of the sample was then pipetted and diluted (1:200) for manual total erythrocyte counts on a hemocytometer, and a 15 ml aliquot was taken for hemoglobin determination with a HemoCueÒ hemoglobin photometer (HemoCue, Inc., Angelholm, Sweden). The remaining sample was centrifuged for analysis of plasma chemistries and electrolytes with an IDEXX VetTestÒ Blood Chemistry Analyzer and VetLyteÒ Electrolyte Analyzer (IDEXX Laboratories, Inc., Westbrook, ME). As many plasma chemistries as possible were analyzed with the available volume using the following hierarchy; total protein, albumin, calcium, phosphorus, uric acid, glucose, triglycerides, aspartate aminotransferase (AST), and electrolytes (Na, K, Cl). These hematology and chemistry parameters had been deemed acceptable for the Western fence lizard based on an in-house analysis comparing blood collection from the post-orbital sinus versus decapitation (unpublished data). All specimens were processed immediately after collection. Due to the extremely small sample volumes (100–200 ml whole blood), samples were utilized in their entirety and dilutions could not be performed for chemistries when initial results exceeded the linearity range of the analyzer. This occurred in six instances for phosphorous and once for triglycerides. For statistical purposes, these data points were entered at their maximum detectable value.
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quantity of Pb. However, lizards did not survive long enough to receive the second dose; these animals died at an effective dose of 4000 mg/kg. Accordingly, animals at the 4000 mg/kg level also died shortly after dosing. Lizards dosed with 3000 mg/kg and higher died the same day they were dosed and showed a rapid response as well, which included limited response to stimulation, brownish coloration and gaping. The female lizard dosed with 2000 mg/kg died 4 days after dosing. The male lizard dosed with 2000 mg/kg was euthanized 6 days after dosing based on lethargy and a lack of eating or drinking in the days following dosing. Lizards that were dosed with 1000 mg/kg or less survived and appeared to recover despite some initial adverse reactions; all were eating crickets 4 days after being dosed. There were no gross observations at necropsy for lizards that survived the 14 days after receiving the one-time dose. 3.3. 14-Day sub-acute toxicity study 3.3.1. Survival There was a dose and time-dependent response of lizards to lead exposure. All lizards dosed with 1000 mg/kg/d died by the end of the study. The average time to death for male and female lizards dosed with 1000 mg/kg/d was 4 and 3 days, respectively. Three females and two males died at the 500 mg/kg/d dose. For males and females exposed to 500 mg/kg/d the average time to death for lizards was 13 and 11 days, respectively. All male and female lizards exposed to lead below 500 mg/kg-bw/d, including controls, survived the 14-day experiment.
No statistical analyses were required for the determination of the approximate lethal dose (ALD). Mean daily cricket consumption, hematological parameters and post-dose behavior were compared using analysis of variance (ANOVA: p < 0.05) provided these data met the assumptions of homogeneity of variance and normality. If data did not meet transformations were performed and mentioned in the results. Lizard body weights were analyzed using a one-way analysis of variance for the ALD and the 14-day sub-acute study; a Mixed-Effects model was used to analyze lizard body weights in the 60-day, sub-chronic study taking into account change in body weight through time.
3.3.2. Growth and cricket consumption There were significant effects of lead on weight gain over the 14-day exposure period (ANOVA p 0.05). Control lizards and those dosed with 31.25 mg/kg/d showed positive weight gain over the 14-day period although lizards at the 31.25 mg/kg/d dose were significantly smaller than control lizards. Lizards dosed with 62.5 mg/kg/d and higher showed weight loss over the 14-day period. There was a significant dose-dependent effect on mean body weight differences with higher doses associated with greater weight loss (p 0.001). The effect was similar in males and females and there were no significant differences between males and females. There were significant effects of lead on cricket consumption (p 0.001). Male and female lizards dosed with 62.5 mg/kg/d and higher ate significantly fewer crickets per day than control lizards and those dosed with 31.25 mg/kg/d. There were no significant effects of sex on cricket consumption.
3. Results
3.4. 60-Day sub-chronic toxicity study
3.1. General procedures
3.4.1. Survival The range of lead exposure levels did not result in widespread mortality; there was a single mortality in both the juvenile and adult lizards at 20 mg/kg/d. Mortality in the juvenile occurred during week 8 of exposure and during week 9 in adults.
2.8. Data analysis
There were no mortalities in control animals during the experiments. In addition, control lizards grew and ate well and generally appeared in good health indicating that husbandry methods were likely adequate for this time period. Importantly, the oral gavage technique proved effective; doses were administered accurately, no regurgitation occurred and lizards were amenable to repeated gavage. The average dose volume for all lizards was 0.07 ml. Dosing averaged approximately 1 min per lizard. Although the technique proved effective, we observed that it was important to apply gentle pressure on the dewlap to avoid bruising or injuring lizards. 3.2. Acute toxicity study The ALD for both male and female lizards was 2000 mg Pb/kg body mass. All mortalities were dose dependent and occurred within one week of dosing. We intended to give lizards at the 8000 mg/kg dose, 2 separate doses 1 h apart to accommodate this
3.4.2. Growth and cricket consumption A Mixed-Effects model was used to analyze changes in lizard body weights through time to account for the fixed effect, Pb dose, and the random effect associated with repeat measurements of individual body weights. There are obvious differences between body weights of juvenile and adult male lizards so the age groups were analyzed separately. Fig. 1a and b shows the change in body weight through time in both juvenile and adult S. occidentalis exposed to inorganic Pb. Inspection of the adult change in body weight shows a clear dose–response relationship although there was not a statistically significant effect of dose only on change in body weight (p ¼ 0.559), the dose time (week) interaction was significant (p 0.001). Inspection of the juvenile body weight data
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Fig. 1. (a) Adult Western fence lizard body weight difference (bwd) through time (week) for different daily doses of Pb (mg/kg/d). Bold line represents no change from initiation of study. Each line in a quadrant represents the body weight changes for and individual lizard for each week of the study. (b) Juvenile Western fence lizard body weight difference (bwd) through time (week) for different daily doses of Pb (mg/kg/d). Bold line represents no change from initiation of study. Each line in a quadrant represents the body weight changes for and individual lizard for each week of the study.
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(Fig. 1.b) does not show as clear a dose response as the adult data although the statistical analysis produced similar results with a significant effect of the dose time interaction (p 0.001). Cricket consumption did not meet the assumptions of ANOVA so the data were analyzed using a Kruskal–Wallis one-way ANOVA on ranks. Data are shown in Fig. 2. There were significant effects of lead on cricket consumption for both juveniles (p < 0.001) and adults (p < 0.001) (Fig. 2). Multiple comparison tests showed that for both juveniles and adults, the 10 and 20 mg/kg/d treatments were significantly different from control while the 1.0 mg/kg/d treatment was not. There was a dose-dependent effect of lead on cricket consumption and, in addition for adults, there appeared to be an impact of time on cricket consumption as well. 3.4.3. Effects of Pb on organ weights There were significant effects of lead on organ weights in both juvenile and adult S. occidentalis. Kidney, brain, testes and body fat were normalized to total body weight before statistical analysis. For normalized testes in juveniles, multiple comparisons (Holm–Sidak) indicated significant differences between the control and 10 (p ¼ 0.033) and 20 (p 0.001) mg/kg/d treatments. Similarly for adults, multiple comparisons indicated a significant difference for normalized testes between control and 10 (p ¼ 0.003) and 20 (p ¼ 0.004) mg/kg/d treatments. Percent body fat was also significantly affected by lead with multiple comparisons showing a significant difference for juveniles between control and the 10 and 20 mg/kg/d treatments (analysis on ranks); for adults there were significant differences between the control and the 10 (p ¼ 0.034) and 20 (p ¼ 0.009) mg/kg/d treatments. Kidney weight showed a positive correlation to lead dose with an increase in normalized kidney weight in juveniles and adults (Fig. 3). Multiple comparison analyses on ranks showed significant differences between control and 10 and 20 mg/kg/d for juveniles (p 0.001) and adults (p 0.001). Only for dark coloration was there a significant effect of age (p 0.002) in addition to an effect of Pb. 3.4.4. Effects of Pb on post-dose behaviors For observations of gaping, arched posture and dark coloration, there were statistically significant effects associated with Pb dose (p < 0.001). Visual inspection confirms a strong positive trend in the frequency of post-dose observations associated with Pb exposure (Fig. 4).
Fig. 3. (a)Effects of daily exposure to different doses of lead for 60 days on adult Western fence lizard organ weights (% body weight) at termination of the study. Error bars are standard deviation. (b) Effects of daily exposure to different doses of lead for 60 days on juvenile Western fence lizard organ weights (% body weights) at termination of the study. Error bars are standard deviation.
3.4.5. Hematological and plasma chemistry results The mean and standard error for all hematology and clinical chemistry parameters are presented in Table 1. The effect of dose
Frequency of Occurrence
0.6 0.5 0.4 0.3
Adult-GAP Juv-GAP Adult-AP Juve-AP Adult-DC Juve-DC
0.2 0.1 0.0 0.0
1.0
10.0
20.0
Dose (mg/kg BW) Fig. 2. Effects of daily exposure of Western fence lizards to different doses of lead on the average number crickets consumed/day over the duration of a 60 d study. Error bars are standard deviation.
Fig. 4. Effects of exposure to lead on post-dose behaviors of Western fence lizards. GAP is gaping, AP is arched posture and DC is dark coloration. Calculated as frequency of occurrence over entire 60 days exposure: number of observations per individual/60 and averaged for given treatment. Error bars are standard deviation.
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Table 1 Hematological Parameters and Chemistry of Western Fence Lizards (Sceloporus occidentalis) exposed to Pb for 60 days. Plasma analyte Albumin (g/dL) AST (U/L)
1.0 mg/kg
10.0 mg/kg
20.0 mg/kg
Adult
Juvenile
Adult
Juvenile
Adult
1.55 0.17 0.85–2.29 (10) 250.1 36.4100.0–405.0 (10) 12.38 0.44 10.90– 13.86 (8) 311.3 37.0245.1–373.0 (3) 13.33 1.39 10.68– 15.41 (3) 4.06 0.30 2.68–5.40 (10) 166.1 25.6 38.5–296.4 (10) 4.1 0.7 1.9–9.5 (10)
1.78 0.13 1.19–2.63 (10) 369.1 47.1139.0– 593.0 (9) 11.68 0.32a 9.83– 12.79 (9) 292.0 9.9266.3–251.9 (8) 13.45 0.73 10.64– 16.10 (8) 4.31 0.14 3.66–4.99 (10) 128.3 31.0 28.2–287.4 (9) 4.8 0.9 1.5–10.5 (9)
1.46 0.21 0.32–2.55 (9) 227.6 23.1153.0– 342.0 (9) 11.63 0.51a 9.66– 13.28 (7) 328.8 22.4255.1– 409.7 (6) 14.39 0.43 12.78– 16.10 (7) 3.72 0.23 2.39–4.94 (10) 163.1 46.2 58.7–485.5 (9) 4.1 0.5 2.1–5.7 (9)
2.14 0.12 1.69–2.79 (10) 352.3 47.6 87.0–592.0 (10) 11.77 0.15 11.12–12.44 (10) 305.8 17.5222.8–404.9 (10) 13.08 0.55 9.93–15.91 (10) 4.70 0.15 4.03–5.41 (10) 122.7 15.6 71.5–231.2 (10) 4.8 0.4 2.8–6.5 (10)
1.27 0.15 0.40–1.83 (10) 165.6 27.7 80.0–292.0 (9) 11.37 0.50 9.04–13.21 (9) 286.4 18.6224.9– 420.3 (9) 13.52 0.76 10.23– 17.18 (9) 3.75 0.18 2.78–4.50 (10) 140.3 20.6 68.3–248.1 (9) 4.3 0.8 2.3–9.2 (9)
1.57 0.23 ND-2.41 (10)
0.75* 0.17 ND-1.69 (9) 1.34 0.15 0.94–2.36 (9) 225.8 55.4 97.0–605.0 163.8 20.3 64.0–251.0 250.1 37.2 52.0–377.0 (10) (9) (9) 10.63 0.36 8.60–12.43 7.93* 0.75 4.36–9.55 9.53* 0.61 6.87–11.34 (10) (7) (9) 247.50 10.46 214.4–323.2 224.2 24.5166.2– 242.9 32.9 15.5–341.8 (10) 314.3 (5) (9) 13.10 0.71 9.51–16.10 12.68 0.73 10.24– 14.13 0.41 12.63– (10) 14.89 (6) 16.10 (9) 4.17 0.32 1.70–5.22 (10) 3.08* 0.24 1.66–4.19 3.82 0.17 3.30–4.95 (9) (9) 224.7 43.4 59.4–406.8 98.36 39.22 ND-240.5 136.2 28.6 ND-235.1 (10) (7) (9) 3.5 0.7 1.4–7.2 (10) 3.5 0.6 1.6–6.6 (8) 4.7 2.0 1.5–20.0 (9)
2.53 0.07 2.17–2.89 (10) 157.9 3.4151.4–164.4 (4) 8.81 (1) 7.71 0.44 6.98–8.84 (4) 125.0 (1) 122.6 2.4116.8–128.4 (4) 28.9 1.7 20.0–35.0 (10) 28.8 1.9 22.0–39.0 (10) 5.5 0.3 3.8–7.1 (10) 5.3 0.2 4.5–6.3 (10)
2.31 0.07 2.07–2.74 (9) 160.5 4.8152.9–173.4 (4) 8.14 0.18 7.62–8.42 (4) 124.1 3.2118.8–132.7 (4) 27.0 1.4 20.0–30.0 (9)
2.57 0.07 2.23–2.96 (10) 162.4 1.6157.8–172.3 (8) 8.72 0.54 5.58–10.54 (8) 125.0 2.7117.4–137.2 (8) 30.4 0.8 26.0–33.0 (10)
2.60 0.11 1.70–2.99 (10) 2.33 0.09 1.66–2.55 (9) 162.9 1.6157.9–170.1 (7) 161.5 4.1153.9–167.9 (3) 6.04 0.59 4.36–8.85 (7) 7.88 1.03 6.13–9.68 (3) 123.89 1.28 119.4–129.3 127.8 2.0124.8–131.7 (7) (3) 24.0 1.9 15.0–35.0 (10) 22.1 2.3 16.0–33.0 (7)
2.48 0.05 2.18–2.72 (9) 165.6 2.3158–174.2 (6) 8.06 0.66 4.78–8.94 (6) 127.8 2.6121.5–139.3 (6) 24.1 2.2 15.0–38.0 (9)
4.9 0.3 3.4–6.7 (9)
5.6 0.3 3.2–6.6 (10)
2.49 0.08 2.11–2.97 (10) 159.5 0.9158.6–160.3 (2) 4.77 0.16 4.61–4.93 (2) 119.6 3.3116.3–122.9 (2) 27.1 2.1 14.0–35.0 (10) 5.0 0.3 3.6–6.1 (10)
5.1 0.3 3.3–6.3 (10)
4.0* 0.2 3.4–4.8 (6)
4.4 0.2 3.8–5.4 (9)
7.1 0.4 4.9–8.6 (10)
6.5 0.3 5.1–7.4 (9)
7.6 0.2 6.2–8.5 (10)
5.6* 0.4 2.9–7.1 (10)
4.0* 0.6 1.6–7.2 (10)
4.2* 0.4 2.7–6.3 (9)
4.0* 0.2 2.9–5.1 (9)
1.30 0.08 1.16–1.43 (3)
1.38 0.05 1.20–1.63 (9) 1.28 0.09 0.86–1.55 (7)
2.51 0.15 1.83–3.11 (10) 161.5 (1)
7.2 0.5 4.8–9.6 (10)
1.18 0.12 0.88–1.44 (4) 1.29 0.01 0.97–1.74 (8)
Juvenile
0.98* 0.08 0.49–1.26 (10) 1.24 0.09 1.08–1.41 (4)
Adult
0.98* 0.08 0.69–1.39 (8)
All values are mean standard error, minimum and maximum, and number of lizards (n) in parentheses. ND ¼ not detected. *Significantly different from control (p 0.05). a Phosphorus levels were >16.10 mg/dL for six lizards and entered at that level for statistical analyses (0.0 mg/kg, n ¼ 2 adults; 1.0 mg/kg, n ¼ 1 juvenile; 10.0 mg/kg, n ¼ 2 adults, and 20.0 mg/kg, n ¼ 1 adult). b Uric acid was >20.0 mg/dL for one adult in the 20 mg/kg treatment and entered at that level for statistical analyses.
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Calcium (mg/dL) Glucose (mg/dL) Phosphorusa (mg/dL) Total Protein (g/dL) Triglycerides (mg/dL) Uric Acidb (mg/dL) Globulin (g/dL) Sodium (mmol/L) Potassium (mmol/L) Chloride (mmol/L) Hematocrit (%) Total Solids (g/dL) Hemoglobin (g/dL) Erythrocytes (106/ml)
0.0 mg/kg Juvenile
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was significant for hematocrit (p ¼ 0.008), total solids (p 0.001), and hemoglobin (p 0.001). Although there was no effect of age alone on hematology, there were age and dose interactions for hemoglobin (p ¼ 0.026) and total RBCs with borderline significance (p ¼ 0.055). In juveniles, hemoglobin was decreased for both the 10 mg/kg/d (p 0.001) and 20 mg/kg/d (p 0.001) treatments with no differences in total RBCs. In adults, hemoglobin (p 0.001, p 0.001) and total erythrocyte count (p ¼ 0.019, p ¼ 0.030) were lower for both the 10 and 20 mg/kg/d treatments, respectively. Thus, with decreased erythrocytes in addition to hemoglobin, the adults appeared more sensitive than the juveniles. The effect of dose was significant for albumin (p 0.001), calcium (p 0.001), glucose (p ¼ 0.003), and total protein (p ¼ 0.005). Albumin, AST, and total protein were analyzed separately for juveniles and adults due to significant effects of age. Calcium in both juveniles (p 0.001) and adults (p 0.001) was significantly lower in the 20 mg/kg/d treatment versus controls. While differences were detected among treatments for potassium, there were no differences compared to controls. Albumin (p ¼ 0.008) and total protein (p ¼ 0.020) in juveniles was lower in the 20 mg/kg/d treatment versus controls, while there were no differences versus controls for AST. 4. Discussion Western fence lizards showed a strong and clear dose-response to Pb exposure under acute and long-term exposures. The approximate lethal dose for both males and females was 2000 mg/kg. Lizards at lower doses (1000 mg/kg and lower) showed some signs of toxicity but appeared to have recovered by the end of the study. Another study aimed at determining the effects of Pb on sprint speed in S. occidentalis showed 30% mortality in lizards dosed with 1000 mg/kg (Holem et al., 2006). While the mortality estimate is not directly comparable to this study because of differences in study design, results from both studies suggest that the LD50 likely lies between 1000 and 2000 mg/kg. Other studies on S. occidentalis have produced LD50 estimates for other xenobiotics using a sequential testing approach (Suski et al., 2008; McFarland et al., 2008) indicating this species is amenable to that design. Generally for the acute and 14-day sub-acute toxicity study, lizards showed sensitivity to Pb similar to birds and mammals. A survey of acute toxicity data for mammals indicates wide variation in sensitivity associated with gavage exposures to Pb. Low-end estimates of the NOAEL range from around 4–10 mg/kg (Junaid et al., 1997; Lorenzo et al., 1978) for mice and rabbits, respectively. In rats, NOAELs have shown a wider range of values from around 200 mg/kg (Petrusz et al., 1979) to above 3000 mg/kg (Holtzman et al., 1982). Based on this study, S. occidentalis would fall near the middle, in terms of sensitivity to Pb in comparison to these toxicity values. The effects of Pb on growth in mammals is similarly variable with NOAELs ranging from 10.7 mg/kg/d for a 10-day gavage exposure in rabbits (Lorenzo et al., 1978) to about 400 mg/kg/d for a 14-day gavage study in rats (Holtzman et al.,1982). The NOAEL for growth in fence lizards exposed to Pb for 14 days of 31.5 mg/kg/d seems comparable, if not somewhat lower than the majority of mammalian NOAELs. Acute toxicity endpoints for avian species exposed to Pb, via oral gavage show mortality ranging from around 125 mg/kg/d for a 35day study on chickens (Vengris and Mare´, 1974) to 625 mg/kg/d for American kestrels (Hoffman et al., 1985). Growth endpoints for avian species, are more similar to the NOAELs generated for fence lizards (31.5 mg/kg/d) with a range of NOAELs for growth from 2.5 to 61.3 mg/kg/d for a variety of species and study durations. There are few actual avian studies where exposure route and duration were similar to those used here. A study on American kestrels
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(Hoffman et al., 1985) where birds were exposed via gavage for 10 days showed a NOAEL for growth of 25.0 mg/kg/d. There were significant effects of Pb on all sub-lethal endpoints evaluated during the 60 d sub-chronic study including cricket consumption, body weight, hematological parameters, clinical chemistry parameters, histopathological observations and postdose behaviors for lizards. There were no strong age-related effects although effects related to body weight seemed greater for adults; for this parameter, growth associated with the juvenile stage likely impacted the response to Pb. While there are some acute and subacute toxicity data for reptiles exposed to Pb, we could not compare the data collected from the present 60 d study with the results of other studies because there are no data on the toxicity of Pb to reptiles for exposures longer than two weeks. In addition, there was a significant interaction between time and dose suggesting that the effects of Pb became more pronounced with duration of exposure; this bears out from visual inspection of the curves. Juveniles clearly grew through the course of the study so any effects on change in body weight are superimposed over changes in body weight associated with somatic growth and maturation. To address the well-recognized shortcomings of using NOAELs and LOAELs (Laskowski, 1995), we conducted a benchmark dose analysis (BMD; Filipsson et al., 2003) to determine the exposure concentration associated with a 10% effect level on body weight. A 10% effect level was chosen because it is thought to be near what might be considered an ecologically meaningful and measurable effect. For adults the exposure level estimated to result in a 10% reduction in body weight was 5.9 mg/kg/d while in juveniles the exposure level was 8.9 mg/kg/d. These values are near the exposure concentration of 10 mg/kg/d which resulted in statistically significant effects on body weight in juvenile and adult lizards. Importantly, the BMD analysis is based on a 60-day exposure although there is an apparent effect of time as well. For shorter duration exposures, it is possible that the BMD for a 10% effect level would be higher. Presently, the U.S. Environmental Protection Agency Ecological Soil Screen Levels (Eco-SSLs) for Pb are 11 mg/kg dry weight in soil for avian wildlife and 56 mg/kg dry weight in soil for mammals (US EPA, 2005a). The results for the BMD analysis for lizards approximate the avian Eco-SSL which, was based on risk to avian ground insectivores (e.g. woodcock). In addition to effects on body weight and food consumption, there were significant effects of Pb on organ weight, hematological parameters and post-dose behaviors. Examination of the organ weight data shows that there were significant effects on kidney weight with a positive relationship to Pb dose and fat and gonad weight, which showed a negative relationship to Pb dose. These results suggest that there may be effects of Pb on the kidney and also that energy acquisition or allocation may have been affected as manifested by a decrease in gonad and body fat weight. Classic signs of Pb induced anemia were observed with decreased hemoglobin at both 10 and 20 mg/kg/d. Hematocrit, total solids, calcium, and glucose were also reduced at 20 mg/kg/ d. Although, adults had decreased RBCs at 10 and 20 mg/kg/d and the juveniles did not, decreases in albumin and total protein were observed in juveniles at 20 mg/kg/d and not in the adults. No effects on hematology or clinical chemistry were observed in the 1.0 mg/kg/d treatment. Post-dose observations indicated that there were dose-related effects on gaping, arched posture and dark coloration. Although the ecological significance of these behaviors is unknown, they warrant further study as alterations in behavior in wild animals may have significant survival consequences. Taken as a whole, there was little observed mortality; however, the sub-lethal effects observed are likely to result in negative effects on survival and reproduction were similar effects to manifest in the field.
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There is a well-recognized lack of data on the toxicity of chemicals to reptilian species (Hopkins, 2000). Research efforts have, for the most part, focused on measuring body burdens in field-collected reptiles, primarily chelonians (Sparling et al., 2000). Importantly, while the toxicity testing system using S. occidentalis seems highly suitable for routine use in developing reptile toxicity data as exemplified by this study and other recently published work (Holem et al., 2006; Suski et al., 2008; McFarland et al., 2008), the applicability of these toxicity data towards other reptilian species remains unknown. Since turtles, snakes and other lizard species can have very different life histories, behaviors, and habitat preferences compared to S. occidentalis, any data generated using this species must be applied with caution when evaluating risks. Nonetheless, recent efforts in reptile toxicity testing (Suski et al., 2008; McFarland et al., 2008) marks an advance in furthering our understanding of the effects of xenobiotics on reptiles by providing a robust, simple, and efficient testing system. Although more reptile toxicity data are becoming available, significant questions remain regarding exposure to reptiles under natural conditions. Future studies regarding the potential risks of contaminants to reptiles should address this source of uncertainty. Exposure studies and estimation methods should reflect the biology of this class of animals with explicit consideration of dermal exposure (many reptiles are in close proximity to the ground), incidental or intentional soil ingestion (Beyer et al., 1994a,b; Rich and Talent, 2009) and bioavailability from soil and food items. Moreover, because reptiles are ectothermic, energy demands and hence consumption rates will be environmentally dependent and exposure estimation methods must account for this fundamental physiological difference compared to birds and mammals. Acknowledgements The authors would like to thank anonymous reviewers for helpful comments. All research was conducted at U.S. Army Center for Health Promotion and Preventive Medicine. This study was conducted consistent with Good Laboratory Practices and through and approved protocol with the Institutional Animal Care and Use Committee. The animal exposures and husbandry were performed in animal facilities fully accredited by the American Association for the Assessment and Accreditation of Laboratory Animal Care. The views expressed in this paper are those of the authors and do no necessarily reflect the views and policies of the U.S. Army. References Beyer, W.N., Connor, E., Gerould, S., 1994a. Survey of soil ingestion by wildlife. The Journal of Wildlife Management 58, 375–382. Beyer, W.N., Connor, E.E., Gerould, S., 1994b. Estimates of soil ingestion by wildlife. The Journal of Wildlife Management 58, 375–382. Brasfield, S.M., Weber, L.P., Talent, L.G., Janz, D.M., 2002. Dose-response and time course relationships for vitellogenin induction in male Western fence lizards (Sceloporus occidentalis) exposed to ethinylestradiol. Environmental Toxicology and Chemistry 21, 1410–1416. Campbell, K.R., Campbell, T.S., 2002. A logical starting point for developing priorities for lizard and snake ecotoxicology: a review of available data. Environmental Toxicology and Chemistry 21, 894–898. Filipsson, A.F., Sand, S., Nilsson, J., Victorin, K., 2003. The benchmark dose methodreview of available models, and recommendations for application in health risk assessment. Critical Reviews in Toxicology 33, 505–542.
Gibbons, J.W., Scott, D.E., Ruan, T.J., Buhlmann, K.A., Tuberville, T.D., Metts, B.S., Greene, J.L., Mills, T., Leiden, Y., Poppy, S., 2000. The global decline of reptiles, De´ja` Vu Amphibians. BioScience 50, 653–666. Hoffman, D.J., Franson, J.C., Pattee, O.H., Bunck, C.M., Anderson, A., 1985. Survival, growth, and accumulation of ingested lead in nestling American Kestrels (Falco sparverius). Archives of Environmental Contamination and Toxicology 14, 89–94. Holem, R.R., Hopkins, W.A., Talent, L.G., 2006. Effect of acute exposure to malathion and lead on sprint performance of the western fence lizard (Sceloporus occidentalis). Archives of Environmental Contamination and Toxicology 51, 111–116. Hopkins, W.A., 2000. Reptile toxicology: opportunities and challenges on the last frontier of vertebrate ecotoxicology. Environmental Toxicology and Chemistry 19, 2391–2393. Holtzman, D., De Vries, C., Nguyen, H., Jameson, N., Olson, J., Carrithers, M., Bensch, K., 1982. Development of resistance to lead encephalopathy during maturation in the rat pup. Journal of Neuropathology and Experimental Neurology 41, 652–663. ITRC (Interstate Technology and Regulatory Council), 2003. Characterization and Remediation of Soils at Closed Small Arms Firing Ranges. Environmental Council of States, Washington, DC, USA. Junaid, M., Chowdhuri, D.K., Narayan, R., Shanker, R., Saxena, D.K., 1997. Leadinduced changes in ovarian follicular development and maturation in mice. Journal of Toxicology and Environmental Health, Part A 50, 31–40. Laskowski, R., 1995. Some good reasons to ban the use of NOEC, LOEC and related concepts in ecotoxicology. Oikos 73, 140–144. Lorenzo, A.V., Gewirtz, M., Averill, D., 1978. CNS lead toxicity in rabbit offspring. Environmental Research 17, 131–150. McFarland, C.A., Quinn Jr., M.J., Bazar, M.A., Remick, A.K., Talent, L.G., Johnson, M.S., 2008. Toxicity of oral exposure to 2,4,6-trinitrotoluence in the Western fence lizard (Sceloporus occidentalis). Environmental Toxicology and Chemistry 27, 1102–1111. Panichayapichet, P., Nitisoravut, S., Simachaya, W., 2007. Spatial distribution and transport of heavy metals in soil, ponded-surface water and grass in a pbcontaminated watershed as to related to land-use practices. Environmental Monitoring and Assessment 135 (1–3), 181–193. Petrusz, P., Weaver, C.M., Grant, L.D., Mushak, P., Krigmann, M.R., 1979. Lead poisoning and reproduction effects on pituitary and serum gonadotropins in neo natal rats. Environmental Research 19 (2), 383–391. Rich, C.N., Talent, L.G., 2009. Soil ingestion may be an important route for the uptake of contaminants by some reptiles. Environmental Toxicology and Chemistry 28 (2), 311–315. Ruth, S.B., 1977. A Comparison of the Demography and Female Reproduction in Sympatric Western Fence Lizards (Sceloporus occidentalis) and Sagebrush Lizards (Sceloporus graciosus) on Mount Diablo, California. PhD thesis. University of California, Berkley, California, USA. Sparling, D.W., Linder, G., Bishop, C.A., 2000. Ecotoxicology of Amphibians and Reptiles. SETAC, Pensacola, FL, USA. Suski, J.G., Salice, C., Houpt, J.T., Bazar, M.A., Talent, L.G., 2008. Dose-related effects following oral exposure of 2,4-dinitrotoluence on the Western fence lizard, Sceloporus occidentalis. Environmental Toxicology and Chemistry 27, 352–359. Talent, L.G., Dumont, J.N., Bantle, J.A., Janz, D.M., Talent, S.G., 2002. Evaluation of Western fence lizards (Sceloporus occidentalis) and Easter fence lizards (Sceloporus undulates) as laboratory reptile models for toxicological investigations. Environmental Toxicology and Chemistry 21, 899–905. Talent, L.G., 2005. Effect of temperature on toxicity of a natural pyrethrin pesticide to green anole lizards (Anolis carolinensis). Environmental Toxicology and Chemistry 24, 3113–3116. US EPA (United States Environmental Protection Agency), 2004. Overview of the Ecological Risk Assessment Process in the Office of Pesticide Programs, U.S. Environmental Protection Agency. Endangered and Threatened Species Effects Determinations. U.S. Environmental Protection Agency, Office of Prevention, Pesticides and Toxic Substances, Office of Pesticide Programs, Washington, DC. January 23, 2004. US EPA (United States Environmental Protection Agency), 2005a. Ecological Soil Screening Levels for Lead. Interim Final – March 2005. OSWER Directive 9285.7– 70 Available from:. http://www.epa.gov/ecotox/ecossl/pdf/eco-ssl_lead.pdf. US EPA (United States Environmental Protection Agency), 2005b. Best Management Practices for Lead at Outdoor Shooting Ranges. EPA-902-B-01–001. Division of Enforcement and Compliance Assistance, New York, NY. Vengris, V.E., Mare´, C.J., 1974. Lead poisoning in chickens and the effect of lead on interferon and antibody production. Canadian Journal of Comparative Medicine 38, 328. Zhang, C., Wu, L., Luo, Y., Zhang, H., Christie, P., 2007. Identifying sources of soil inorganic pollutants on a regional scale using a multivariate statistical approach: role of pollutant migration and soil physiochemical properties. Environmental Pollution 151, 470–476.
Contents lists available at ScienceDirect
Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Effects of inorganic lead on Western fence lizards (Sceloporus occidentalis) Christopher J. Salice a, *, Jamie G. Suski a,1, Matthew A. Bazar a, Larry G. Talent b a b
US Army Center for Health Promotion and Preventive Medicine, Aberdeen Proving Ground, MD 21010, USA Oklahoma State University, Department of Natural Resource Ecology and Management, Stillwater, OK 74078, USA
The Western fence lizard, Sceloporus occidentalis, is sensitive to Pb and is a useful laboratory model for ecotoxicological testing of reptiles.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 30 January 2009 Received in revised form 3 June 2009 Accepted 15 June 2009
Although anthropogenic pollutants are thought to threaten reptilian species, there are few toxicity studies on reptiles. We evaluated the toxicity of Pb as lead acetate to the Western fence lizard (Sceloporus occidentalis). The acute lethal dose and sub-acute (14-day) toxicity studies were used to narrow exposure concentrations for a sub-chronic (60-day) study. In the sub-chronic study, adult and juvenile male lizards were dosed via gavage with 0, 1, 10 and 20 mg Pb/kg-bw/day. Mortality was limited and occurred only at the highest dose (20 mg Pb/kg-bw/d). There were statistically significant sub-lethal effects of 10 and 20 mg Pb/kg-bw/d on body weight, cricket consumption, organ weight, hematological parameters and post-dose behaviors. Of these, Pb-induced changes in body weight are most useful for ecological risk assessment because it is linked to fitness in wild lizard populations. The Western fence lizard is a useful model for reptilian toxicity studies. Ó 2009 Elsevier Ltd. All rights reserved.
Keywords: Reptiles Lizard Lead Toxicity Metals
1. Introduction An important objective in ecotoxicology is to generate data that can be used to facilitate environmental management, principally through the ecological risk assessment (ERA) process. Frequently, however, risk assessments must be conducted with incomplete toxicological datasets on receptors of interest and, hence, rely on toxicity data for surrogate species which, can introduce significant uncertainty. For example, U.S. EPA guidance on risk assessment of pesticides suggests that birds can be used as surrogates for reptiles and that fish can be used as surrogate taxa for amphibians (US EPA, 2004). Although risk estimates based on surrogate species are convenient, they may lead to significant errors in projecting risk; risk estimates may be too high or too low resulting in cost-prohibitive overprotection and environmentally damaging under-protection, respectively. At a minimum it would be beneficial to have laboratory models representing each vertebrate class. Ideally, the availability of more than one representative for each class would provide a better perspective on the sensitivity of a given class to a toxicant. While
* Corresponding author. Present address: Texas Tech University, The Institute of Environmental and Human Health, Box 41163, USA. Tel.: þ1 806 885 4567x229; fax: þ1 806 885 2132. E-mail addresses: [email protected] (C.J. Salice), [email protected] (J.G. Suski), [email protected] (M.A. Bazar), [email protected] (L.G. Talent). 1 Present address: Texas Tech University, Department of Biology. 0269-7491/$ – see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2009.06.013
there are standard toxicity test species for most animal classes, to date there have been no reptilian species routinely used for toxicity testing. The animal class, Reptilia, has been historically under-represented by the field of toxicology (Hopkins, 2000; Campbell and Campbell, 2002). Despite the fact that reptiles are widespread, there has been relatively few laboratory toxicity studies conducted. Reptiles may have been excluded from consideration because it was believed that benchmarks protective of mammalian and avian receptors were likely also protective of reptilian receptors. However, the approach to using non-reptile toxicity data was necessitated, in part, by the lack of a viable laboratory reptile model, which was likely driven by the fact that many reptiles can be slow to reach reproductive maturity and may not be conducive to convenient laboratory rearing due to size or special husbandry considerations. Also, it is possible that while methods and concepts related to ecotoxicology and ERAs were developing, focus was placed on more prominent and readily available mammalian and avian models. There is, however, a growing interest in assessing risks to historically under-represented species such as those belonging to the class reptilia (Hopkins, 2000; Campbell and Campbell, 2002). In addition, there is evidence that lizard populations are declining and of the six main threats discussed by Gibbons et al. (2000), anthropogenic pollutants are the greatest. The development of a routine reptilian toxicity-testing model could foster the development of reptilian toxicity benchmarks for use in ERAs and ultimately contribute to improved management decisions with regard to naturally occurring reptile populations.
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Previous efforts to develop a reptile toxicity model resulted in the identification of several populations of Eastern (Sceloporus undulatus) and Western (Sceloporus occidentalis) fence lizards that can be reared and bred successfully in the laboratory (Talent et al., 2002). Subsequently, Western fence lizards have been used to evaluate the toxicity of several toxicants including military compounds (Suski et al., 2008; McFarland et al., 2008), pesticides (Talent, 2005; Holem et al., 2006) and endocrine disrupters (Brasfield et al., 2002). The objective of the present study was to build upon previous efforts to further utilize a lizard model for routine toxicity testing. The approach was modeled after standardized mammalian toxicity tests that progress in a hierarchical fashion from acute lethal dose, short-term sub-acute exposures, and longer sub-chronic exposures (Suski et al., 2008). Lead (Pb) is a naturally occurring metal but is not considered an essential nutrient for plants or animals (US EPA, 2005a). It is widely recognized as an important environmental contaminant from industrial activities such as smelting and mining (Zhang et al., 2007; Panichayapichet et al., 2007), and on Small Arms Firing Ranges (SAFRs) from the use of lead shot (ITRC, 2003; US EPA, 2005b). Since it is likely that reptiles inhabiting these areas would be exposed to Pb, reptilian toxicity data could be used in screeninglevel or refined ERAs. Hence, the main objectives of the present study were to evaluate the acute and sub-chronic toxicity of lead to S. occidentalis and to further vet the testing system for potential development as a routine reptilian toxicity bioassay. Acute (single dose), sub-acute (daily dose for 14 days) and sub-chronic (daily dose for 60 days) studies were conducted. In addition the subchronic study was conducted for both juvenile and adult male Western fence lizards to obtain some insight into the potential for age-specific sensitivity that may occur over longer-term exposures. 2. Materials and methods 2.1. Lizards and husbandry Adult (male and female) and juvenile (male) S. occidentalis were obtained from laboratory-reared stocks maintained at Oklahoma State University that were descendents of wild-caught animals from the San Joaquin Valley (CA, USA). The approximate age of lizards was 4 months for juvenile and one year for adults. All lizards were maintained in the laboratory for two weeks prior to study initiation. Pre-study conditions were the same as conditions during the study. General husbandry information for S. occidentalis has been previously published (Talent et al., 2002; Suski et al., 2008; McFarland et al., 2008). Briefly, lizards were maintained individually in modified, clear-plastic mouse cages. The cage dimensions were 20 cm deep, 20 cm wide and 40.5 cm long. Wood chip bedding, approximately 1.3 cm deep, a 10 cm PVC (3.8 cm diameter) and a water dish (small plastic petri dish) were placed in every cage. Water was changed daily, woodchips and cages were changed once per week. Lizards were fed 1–4 house crickets (Acheta domesticus) per day depending on the size of the cricket; lizards on study were given 2, 1.3 cm crickets per day. Crickets were dusted with Rep-CalÒ (Rep-Cal Research Labs, Los Gatos, CA, USA), a vitamin and mineral supplement to help prevent metabolic bone disease. Importantly, many reptiles require a temperature gradient for optimal health. A temperature gradient from 22 to 34 C was maintained in each cage by keeping room temperature at 22 C and using an electric heat strip to warm one end of the cage. Light was ambient fluorescent supplemented with incandescent both on a 12:12 light:dark cycle. Humidity ranged between 30 and 70%. This study was conducted under a Good Laboratory Practices (GLP) protocol approved by United States Army Center for Health Promotion and Preventive Medicine (USACHPPM) Animal Care and Use Committee (IACUC). As with any toxicity study, the quality of results is strongly dependent on accurate dosing technique. The dosing method employed allowed oral dosing of adult lizards using a 1 cc syringe and a 22 gauge stainless steel feeding tube; a pipette may also be used (Suski et al., 2008). Doses were delivered orally because lizards are likely to be exposed to many toxicants via the diet. In addition, oral delivery was a more precise dosing regimen compared to feeding studies. Dosing was found to be more efficient with two people. Standard dosing procedure involved one person obtaining the lizard and while holding it in one hand, gently applying steady downward pressure on the dewlap until the mouth opens. The person administering the dose gently places one end of a small wooden dowel in the corner of the mouth to ensure the mouth does not close. The dose is administered to the rear of the mouth but not past the throat. Unlike in many rodent studies, the dosing
tube was not placed directly into the stomach as this was thought to be too traumatic for lizards. While a pilot study indicated that adult lizards could be dosed with as much as 0.15 ml, we found that dosing success was nearly 100% when dosing volume was between 0.05 and 0.10 ml. 2.2. Test article Lead acetate was obtained from Aldrich chemical company. Because Lead acetate is soluble, distilled water was used as the carrier for dosing experiments; analytical results showed that Pb concentrations in control water would result in doses to lizards of about 0.008 mg/kg, well below the lowest dose tested of 1.0 mg/ kg. All dosing solutions were analyzed for total lead concentration following EPA method 6010B (ICP) using a Perkin Elmer Spectrometer (Optima 3000 XL). Test concentrations were verified prior to and after the experiment. 2.3. Experimental design The approach to developing the Western fence lizard as a viable toxicity-testing model was based largely on standardized approaches for mammalian testing (Suski et al., 2008; McFarland et al., 2008). In the present study, a hierarchy of studies was used to proceed from acute to sub-chronic studies. This methodical approach helped to focus, for a given study duration, the range of toxicant concentrations needed to generate dose-dependent variation in response. In addition, the method ultimately reduced the total number of animals required. 2.4. Acute toxicity study The first step in evaluating the toxicity of lead acetate to lizards was to determine an Approximate Lethal Dose (ALD) of Pb, which is the lowest dose that causes mortality within 14 days following a single administered dose. The method requires using only one animal of each sex at each dose level. Although this method does not result in a statistical estimate of a lethal dose such as an LD50, it approximates the lethal dose while reducing the total number of animals used; the ALD establishes the upper bound for longer duration toxicity studies. Although the primary endpoint of interest in the ALD was mortality, other endpoints were evaluated such as growth, food consumption, behavior and some hematological parameters. Average mass of lizards used for the ALD was 11.3 þ/ 1.14 g for males and 11.9 þ/ 0.87 g for females. All lizards were adults. One lizard of each sex was given a deionized water control, or as lead acetate solutions with one of the following doses of Pb: 100, 200, 500, 1000, 2000, 3000, 4000 and 8000 mg Pb/kg body mass, denoted as mg/kg hereafter. To achieve a dose of 8000 mg Pb/kg body mass, two separate dosing events, separated by 1 h, were required. Pilot studies indicated that this was sufficient time to allow 2 doses near 0.15 ml to be administered and it was unlikely there would be any time-delay effects on toxicity. 2.5. 14-Day sub-acute toxicity study The 14-day study was designed as a range-finding study to further define suitable doses of Pb for use in the longer-term sub-chronic study. Six adult lizards per sex were used for each treatment level and given one dose per day for 14 consecutive days. The doses were 0 (water control), 31.25, 62.50, 125, 250, 500, and 1000 mg Pb/kg body mass/d, denoted as mg/kg/d hereafter. Six adult animals of each sex per dose provided enough insight into potential levels of toxicity while also minimizing the number of animals used. Lizards were observed daily for signs of toxicity including morbidity, ataxia or hyperactivity and apparent dosing-related behaviors (explained below). Each lizard was given 2–4 fresh crickets each day and the number of crickets remaining on the following day was recorded as a measure of food consumption. Lizards were weighed before and during the 14-day study on day 3, 1, 0, 7 and 14. Lizards were sorted into treatment groups using LabcatÒ such that there were no significant differences in weight among groups. Treatments were then randomly assigned to each group. All lizards alive on day 14 were euthanized and necropsied. Whole blood was collected with heparinized capillary tubes for hematology and clinical chemistry analyses. 2.6. 60-Day sub-chronic toxicity study The length of the sub-chronic exposure was 60 days, which is approximately 3–16% of the life span of S. occidentalis (Ruth, 1977). Forty juvenile and forty adult male lizards were randomly sorted into four dose groups. Female lizards were excluded from the sub-chronic toxicity study because they can spontaneously enter vitellogenesis and internal egg masses can account for 20% of body mass, potentially confounding dosing. Based on the results of the 14-day study, doses were selected as 0, 1, 10, or 20 mg/ kg/d. Doses were administered to each lizard, every day for 60 consecutive days. All animals were weighed on day 3, 1, 0, 3, 7 and weekly thereafter. Lizards were fed 2 crickets per day and were allowed 24 h to consume crickets. All uneaten crickets were removed the following day and counted to monitor food consumption. Daily observations were made on condition; if animals appeared lethargic or moribund or did not respond to stimulus or righting, they were euthanized. Any animals
C.J. Salice et al. / Environmental Pollution 157 (2009) 3457–3464 remaining alive on day 60 were euthanized using CO2 followed by decapitation. Whole blood was collected with heparinized capillary tubes for hematological and clinical chemistry analyses. Additionally, brain, liver, heart, kidneys and gonads were removed, weighed and preserved in formalin. Post-dose observations included turning dark in coloration, assuming arched posture and/or mouth gaping were recorded during the 60-d study. Dark coloration could be described as an overall darkening of the skin (from gray to charcoal gray/ black) accompanied by sharply contrasting lighter markings that are less pronounced under normal conditions. Arched posture describes lizards that appear to extend the forelimbs so that the head is facing nearly vertical with the hind limbs and tail resting on the ground. This behavior is different than head-bobbing or push-ups in that there was no cyclical movement; once forelimbs were extended, lizards remained in that position. Gaping was characterized by either, keeping the mouth open or repeated opening and closing. All lizards were observed for occurrence of dark coloration, arched posture or gaping within 30 min after a dose had been administered. 2.7. Hematology and clinical chemistries Following euthanasia, blood samples were collected with several heparinized capillary tubes (lithium heparin). The heparinized sample was then discharged from each capillary tube into a plain 1.5 ml microtube for analysis. Approximately 35 ml of the whole blood sample was drawn off with a plain capillary tube and centrifuged for manual determination of hematocrit and total solids with a card reader and refractometer. 50 ml of the sample was then pipetted and diluted (1:200) for manual total erythrocyte counts on a hemocytometer, and a 15 ml aliquot was taken for hemoglobin determination with a HemoCueÒ hemoglobin photometer (HemoCue, Inc., Angelholm, Sweden). The remaining sample was centrifuged for analysis of plasma chemistries and electrolytes with an IDEXX VetTestÒ Blood Chemistry Analyzer and VetLyteÒ Electrolyte Analyzer (IDEXX Laboratories, Inc., Westbrook, ME). As many plasma chemistries as possible were analyzed with the available volume using the following hierarchy; total protein, albumin, calcium, phosphorus, uric acid, glucose, triglycerides, aspartate aminotransferase (AST), and electrolytes (Na, K, Cl). These hematology and chemistry parameters had been deemed acceptable for the Western fence lizard based on an in-house analysis comparing blood collection from the post-orbital sinus versus decapitation (unpublished data). All specimens were processed immediately after collection. Due to the extremely small sample volumes (100–200 ml whole blood), samples were utilized in their entirety and dilutions could not be performed for chemistries when initial results exceeded the linearity range of the analyzer. This occurred in six instances for phosphorous and once for triglycerides. For statistical purposes, these data points were entered at their maximum detectable value.
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quantity of Pb. However, lizards did not survive long enough to receive the second dose; these animals died at an effective dose of 4000 mg/kg. Accordingly, animals at the 4000 mg/kg level also died shortly after dosing. Lizards dosed with 3000 mg/kg and higher died the same day they were dosed and showed a rapid response as well, which included limited response to stimulation, brownish coloration and gaping. The female lizard dosed with 2000 mg/kg died 4 days after dosing. The male lizard dosed with 2000 mg/kg was euthanized 6 days after dosing based on lethargy and a lack of eating or drinking in the days following dosing. Lizards that were dosed with 1000 mg/kg or less survived and appeared to recover despite some initial adverse reactions; all were eating crickets 4 days after being dosed. There were no gross observations at necropsy for lizards that survived the 14 days after receiving the one-time dose. 3.3. 14-Day sub-acute toxicity study 3.3.1. Survival There was a dose and time-dependent response of lizards to lead exposure. All lizards dosed with 1000 mg/kg/d died by the end of the study. The average time to death for male and female lizards dosed with 1000 mg/kg/d was 4 and 3 days, respectively. Three females and two males died at the 500 mg/kg/d dose. For males and females exposed to 500 mg/kg/d the average time to death for lizards was 13 and 11 days, respectively. All male and female lizards exposed to lead below 500 mg/kg-bw/d, including controls, survived the 14-day experiment.
No statistical analyses were required for the determination of the approximate lethal dose (ALD). Mean daily cricket consumption, hematological parameters and post-dose behavior were compared using analysis of variance (ANOVA: p < 0.05) provided these data met the assumptions of homogeneity of variance and normality. If data did not meet transformations were performed and mentioned in the results. Lizard body weights were analyzed using a one-way analysis of variance for the ALD and the 14-day sub-acute study; a Mixed-Effects model was used to analyze lizard body weights in the 60-day, sub-chronic study taking into account change in body weight through time.
3.3.2. Growth and cricket consumption There were significant effects of lead on weight gain over the 14-day exposure period (ANOVA p 0.05). Control lizards and those dosed with 31.25 mg/kg/d showed positive weight gain over the 14-day period although lizards at the 31.25 mg/kg/d dose were significantly smaller than control lizards. Lizards dosed with 62.5 mg/kg/d and higher showed weight loss over the 14-day period. There was a significant dose-dependent effect on mean body weight differences with higher doses associated with greater weight loss (p 0.001). The effect was similar in males and females and there were no significant differences between males and females. There were significant effects of lead on cricket consumption (p 0.001). Male and female lizards dosed with 62.5 mg/kg/d and higher ate significantly fewer crickets per day than control lizards and those dosed with 31.25 mg/kg/d. There were no significant effects of sex on cricket consumption.
3. Results
3.4. 60-Day sub-chronic toxicity study
3.1. General procedures
3.4.1. Survival The range of lead exposure levels did not result in widespread mortality; there was a single mortality in both the juvenile and adult lizards at 20 mg/kg/d. Mortality in the juvenile occurred during week 8 of exposure and during week 9 in adults.
2.8. Data analysis
There were no mortalities in control animals during the experiments. In addition, control lizards grew and ate well and generally appeared in good health indicating that husbandry methods were likely adequate for this time period. Importantly, the oral gavage technique proved effective; doses were administered accurately, no regurgitation occurred and lizards were amenable to repeated gavage. The average dose volume for all lizards was 0.07 ml. Dosing averaged approximately 1 min per lizard. Although the technique proved effective, we observed that it was important to apply gentle pressure on the dewlap to avoid bruising or injuring lizards. 3.2. Acute toxicity study The ALD for both male and female lizards was 2000 mg Pb/kg body mass. All mortalities were dose dependent and occurred within one week of dosing. We intended to give lizards at the 8000 mg/kg dose, 2 separate doses 1 h apart to accommodate this
3.4.2. Growth and cricket consumption A Mixed-Effects model was used to analyze changes in lizard body weights through time to account for the fixed effect, Pb dose, and the random effect associated with repeat measurements of individual body weights. There are obvious differences between body weights of juvenile and adult male lizards so the age groups were analyzed separately. Fig. 1a and b shows the change in body weight through time in both juvenile and adult S. occidentalis exposed to inorganic Pb. Inspection of the adult change in body weight shows a clear dose–response relationship although there was not a statistically significant effect of dose only on change in body weight (p ¼ 0.559), the dose time (week) interaction was significant (p 0.001). Inspection of the juvenile body weight data
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Fig. 1. (a) Adult Western fence lizard body weight difference (bwd) through time (week) for different daily doses of Pb (mg/kg/d). Bold line represents no change from initiation of study. Each line in a quadrant represents the body weight changes for and individual lizard for each week of the study. (b) Juvenile Western fence lizard body weight difference (bwd) through time (week) for different daily doses of Pb (mg/kg/d). Bold line represents no change from initiation of study. Each line in a quadrant represents the body weight changes for and individual lizard for each week of the study.
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(Fig. 1.b) does not show as clear a dose response as the adult data although the statistical analysis produced similar results with a significant effect of the dose time interaction (p 0.001). Cricket consumption did not meet the assumptions of ANOVA so the data were analyzed using a Kruskal–Wallis one-way ANOVA on ranks. Data are shown in Fig. 2. There were significant effects of lead on cricket consumption for both juveniles (p < 0.001) and adults (p < 0.001) (Fig. 2). Multiple comparison tests showed that for both juveniles and adults, the 10 and 20 mg/kg/d treatments were significantly different from control while the 1.0 mg/kg/d treatment was not. There was a dose-dependent effect of lead on cricket consumption and, in addition for adults, there appeared to be an impact of time on cricket consumption as well. 3.4.3. Effects of Pb on organ weights There were significant effects of lead on organ weights in both juvenile and adult S. occidentalis. Kidney, brain, testes and body fat were normalized to total body weight before statistical analysis. For normalized testes in juveniles, multiple comparisons (Holm–Sidak) indicated significant differences between the control and 10 (p ¼ 0.033) and 20 (p 0.001) mg/kg/d treatments. Similarly for adults, multiple comparisons indicated a significant difference for normalized testes between control and 10 (p ¼ 0.003) and 20 (p ¼ 0.004) mg/kg/d treatments. Percent body fat was also significantly affected by lead with multiple comparisons showing a significant difference for juveniles between control and the 10 and 20 mg/kg/d treatments (analysis on ranks); for adults there were significant differences between the control and the 10 (p ¼ 0.034) and 20 (p ¼ 0.009) mg/kg/d treatments. Kidney weight showed a positive correlation to lead dose with an increase in normalized kidney weight in juveniles and adults (Fig. 3). Multiple comparison analyses on ranks showed significant differences between control and 10 and 20 mg/kg/d for juveniles (p 0.001) and adults (p 0.001). Only for dark coloration was there a significant effect of age (p 0.002) in addition to an effect of Pb. 3.4.4. Effects of Pb on post-dose behaviors For observations of gaping, arched posture and dark coloration, there were statistically significant effects associated with Pb dose (p < 0.001). Visual inspection confirms a strong positive trend in the frequency of post-dose observations associated with Pb exposure (Fig. 4).
Fig. 3. (a)Effects of daily exposure to different doses of lead for 60 days on adult Western fence lizard organ weights (% body weight) at termination of the study. Error bars are standard deviation. (b) Effects of daily exposure to different doses of lead for 60 days on juvenile Western fence lizard organ weights (% body weights) at termination of the study. Error bars are standard deviation.
3.4.5. Hematological and plasma chemistry results The mean and standard error for all hematology and clinical chemistry parameters are presented in Table 1. The effect of dose
Frequency of Occurrence
0.6 0.5 0.4 0.3
Adult-GAP Juv-GAP Adult-AP Juve-AP Adult-DC Juve-DC
0.2 0.1 0.0 0.0
1.0
10.0
20.0
Dose (mg/kg BW) Fig. 2. Effects of daily exposure of Western fence lizards to different doses of lead on the average number crickets consumed/day over the duration of a 60 d study. Error bars are standard deviation.
Fig. 4. Effects of exposure to lead on post-dose behaviors of Western fence lizards. GAP is gaping, AP is arched posture and DC is dark coloration. Calculated as frequency of occurrence over entire 60 days exposure: number of observations per individual/60 and averaged for given treatment. Error bars are standard deviation.
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Table 1 Hematological Parameters and Chemistry of Western Fence Lizards (Sceloporus occidentalis) exposed to Pb for 60 days. Plasma analyte Albumin (g/dL) AST (U/L)
1.0 mg/kg
10.0 mg/kg
20.0 mg/kg
Adult
Juvenile
Adult
Juvenile
Adult
1.55 0.17 0.85–2.29 (10) 250.1 36.4100.0–405.0 (10) 12.38 0.44 10.90– 13.86 (8) 311.3 37.0245.1–373.0 (3) 13.33 1.39 10.68– 15.41 (3) 4.06 0.30 2.68–5.40 (10) 166.1 25.6 38.5–296.4 (10) 4.1 0.7 1.9–9.5 (10)
1.78 0.13 1.19–2.63 (10) 369.1 47.1139.0– 593.0 (9) 11.68 0.32a 9.83– 12.79 (9) 292.0 9.9266.3–251.9 (8) 13.45 0.73 10.64– 16.10 (8) 4.31 0.14 3.66–4.99 (10) 128.3 31.0 28.2–287.4 (9) 4.8 0.9 1.5–10.5 (9)
1.46 0.21 0.32–2.55 (9) 227.6 23.1153.0– 342.0 (9) 11.63 0.51a 9.66– 13.28 (7) 328.8 22.4255.1– 409.7 (6) 14.39 0.43 12.78– 16.10 (7) 3.72 0.23 2.39–4.94 (10) 163.1 46.2 58.7–485.5 (9) 4.1 0.5 2.1–5.7 (9)
2.14 0.12 1.69–2.79 (10) 352.3 47.6 87.0–592.0 (10) 11.77 0.15 11.12–12.44 (10) 305.8 17.5222.8–404.9 (10) 13.08 0.55 9.93–15.91 (10) 4.70 0.15 4.03–5.41 (10) 122.7 15.6 71.5–231.2 (10) 4.8 0.4 2.8–6.5 (10)
1.27 0.15 0.40–1.83 (10) 165.6 27.7 80.0–292.0 (9) 11.37 0.50 9.04–13.21 (9) 286.4 18.6224.9– 420.3 (9) 13.52 0.76 10.23– 17.18 (9) 3.75 0.18 2.78–4.50 (10) 140.3 20.6 68.3–248.1 (9) 4.3 0.8 2.3–9.2 (9)
1.57 0.23 ND-2.41 (10)
0.75* 0.17 ND-1.69 (9) 1.34 0.15 0.94–2.36 (9) 225.8 55.4 97.0–605.0 163.8 20.3 64.0–251.0 250.1 37.2 52.0–377.0 (10) (9) (9) 10.63 0.36 8.60–12.43 7.93* 0.75 4.36–9.55 9.53* 0.61 6.87–11.34 (10) (7) (9) 247.50 10.46 214.4–323.2 224.2 24.5166.2– 242.9 32.9 15.5–341.8 (10) 314.3 (5) (9) 13.10 0.71 9.51–16.10 12.68 0.73 10.24– 14.13 0.41 12.63– (10) 14.89 (6) 16.10 (9) 4.17 0.32 1.70–5.22 (10) 3.08* 0.24 1.66–4.19 3.82 0.17 3.30–4.95 (9) (9) 224.7 43.4 59.4–406.8 98.36 39.22 ND-240.5 136.2 28.6 ND-235.1 (10) (7) (9) 3.5 0.7 1.4–7.2 (10) 3.5 0.6 1.6–6.6 (8) 4.7 2.0 1.5–20.0 (9)
2.53 0.07 2.17–2.89 (10) 157.9 3.4151.4–164.4 (4) 8.81 (1) 7.71 0.44 6.98–8.84 (4) 125.0 (1) 122.6 2.4116.8–128.4 (4) 28.9 1.7 20.0–35.0 (10) 28.8 1.9 22.0–39.0 (10) 5.5 0.3 3.8–7.1 (10) 5.3 0.2 4.5–6.3 (10)
2.31 0.07 2.07–2.74 (9) 160.5 4.8152.9–173.4 (4) 8.14 0.18 7.62–8.42 (4) 124.1 3.2118.8–132.7 (4) 27.0 1.4 20.0–30.0 (9)
2.57 0.07 2.23–2.96 (10) 162.4 1.6157.8–172.3 (8) 8.72 0.54 5.58–10.54 (8) 125.0 2.7117.4–137.2 (8) 30.4 0.8 26.0–33.0 (10)
2.60 0.11 1.70–2.99 (10) 2.33 0.09 1.66–2.55 (9) 162.9 1.6157.9–170.1 (7) 161.5 4.1153.9–167.9 (3) 6.04 0.59 4.36–8.85 (7) 7.88 1.03 6.13–9.68 (3) 123.89 1.28 119.4–129.3 127.8 2.0124.8–131.7 (7) (3) 24.0 1.9 15.0–35.0 (10) 22.1 2.3 16.0–33.0 (7)
2.48 0.05 2.18–2.72 (9) 165.6 2.3158–174.2 (6) 8.06 0.66 4.78–8.94 (6) 127.8 2.6121.5–139.3 (6) 24.1 2.2 15.0–38.0 (9)
4.9 0.3 3.4–6.7 (9)
5.6 0.3 3.2–6.6 (10)
2.49 0.08 2.11–2.97 (10) 159.5 0.9158.6–160.3 (2) 4.77 0.16 4.61–4.93 (2) 119.6 3.3116.3–122.9 (2) 27.1 2.1 14.0–35.0 (10) 5.0 0.3 3.6–6.1 (10)
5.1 0.3 3.3–6.3 (10)
4.0* 0.2 3.4–4.8 (6)
4.4 0.2 3.8–5.4 (9)
7.1 0.4 4.9–8.6 (10)
6.5 0.3 5.1–7.4 (9)
7.6 0.2 6.2–8.5 (10)
5.6* 0.4 2.9–7.1 (10)
4.0* 0.6 1.6–7.2 (10)
4.2* 0.4 2.7–6.3 (9)
4.0* 0.2 2.9–5.1 (9)
1.30 0.08 1.16–1.43 (3)
1.38 0.05 1.20–1.63 (9) 1.28 0.09 0.86–1.55 (7)
2.51 0.15 1.83–3.11 (10) 161.5 (1)
7.2 0.5 4.8–9.6 (10)
1.18 0.12 0.88–1.44 (4) 1.29 0.01 0.97–1.74 (8)
Juvenile
0.98* 0.08 0.49–1.26 (10) 1.24 0.09 1.08–1.41 (4)
Adult
0.98* 0.08 0.69–1.39 (8)
All values are mean standard error, minimum and maximum, and number of lizards (n) in parentheses. ND ¼ not detected. *Significantly different from control (p 0.05). a Phosphorus levels were >16.10 mg/dL for six lizards and entered at that level for statistical analyses (0.0 mg/kg, n ¼ 2 adults; 1.0 mg/kg, n ¼ 1 juvenile; 10.0 mg/kg, n ¼ 2 adults, and 20.0 mg/kg, n ¼ 1 adult). b Uric acid was >20.0 mg/dL for one adult in the 20 mg/kg treatment and entered at that level for statistical analyses.
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Calcium (mg/dL) Glucose (mg/dL) Phosphorusa (mg/dL) Total Protein (g/dL) Triglycerides (mg/dL) Uric Acidb (mg/dL) Globulin (g/dL) Sodium (mmol/L) Potassium (mmol/L) Chloride (mmol/L) Hematocrit (%) Total Solids (g/dL) Hemoglobin (g/dL) Erythrocytes (106/ml)
0.0 mg/kg Juvenile
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was significant for hematocrit (p ¼ 0.008), total solids (p 0.001), and hemoglobin (p 0.001). Although there was no effect of age alone on hematology, there were age and dose interactions for hemoglobin (p ¼ 0.026) and total RBCs with borderline significance (p ¼ 0.055). In juveniles, hemoglobin was decreased for both the 10 mg/kg/d (p 0.001) and 20 mg/kg/d (p 0.001) treatments with no differences in total RBCs. In adults, hemoglobin (p 0.001, p 0.001) and total erythrocyte count (p ¼ 0.019, p ¼ 0.030) were lower for both the 10 and 20 mg/kg/d treatments, respectively. Thus, with decreased erythrocytes in addition to hemoglobin, the adults appeared more sensitive than the juveniles. The effect of dose was significant for albumin (p 0.001), calcium (p 0.001), glucose (p ¼ 0.003), and total protein (p ¼ 0.005). Albumin, AST, and total protein were analyzed separately for juveniles and adults due to significant effects of age. Calcium in both juveniles (p 0.001) and adults (p 0.001) was significantly lower in the 20 mg/kg/d treatment versus controls. While differences were detected among treatments for potassium, there were no differences compared to controls. Albumin (p ¼ 0.008) and total protein (p ¼ 0.020) in juveniles was lower in the 20 mg/kg/d treatment versus controls, while there were no differences versus controls for AST. 4. Discussion Western fence lizards showed a strong and clear dose-response to Pb exposure under acute and long-term exposures. The approximate lethal dose for both males and females was 2000 mg/kg. Lizards at lower doses (1000 mg/kg and lower) showed some signs of toxicity but appeared to have recovered by the end of the study. Another study aimed at determining the effects of Pb on sprint speed in S. occidentalis showed 30% mortality in lizards dosed with 1000 mg/kg (Holem et al., 2006). While the mortality estimate is not directly comparable to this study because of differences in study design, results from both studies suggest that the LD50 likely lies between 1000 and 2000 mg/kg. Other studies on S. occidentalis have produced LD50 estimates for other xenobiotics using a sequential testing approach (Suski et al., 2008; McFarland et al., 2008) indicating this species is amenable to that design. Generally for the acute and 14-day sub-acute toxicity study, lizards showed sensitivity to Pb similar to birds and mammals. A survey of acute toxicity data for mammals indicates wide variation in sensitivity associated with gavage exposures to Pb. Low-end estimates of the NOAEL range from around 4–10 mg/kg (Junaid et al., 1997; Lorenzo et al., 1978) for mice and rabbits, respectively. In rats, NOAELs have shown a wider range of values from around 200 mg/kg (Petrusz et al., 1979) to above 3000 mg/kg (Holtzman et al., 1982). Based on this study, S. occidentalis would fall near the middle, in terms of sensitivity to Pb in comparison to these toxicity values. The effects of Pb on growth in mammals is similarly variable with NOAELs ranging from 10.7 mg/kg/d for a 10-day gavage exposure in rabbits (Lorenzo et al., 1978) to about 400 mg/kg/d for a 14-day gavage study in rats (Holtzman et al.,1982). The NOAEL for growth in fence lizards exposed to Pb for 14 days of 31.5 mg/kg/d seems comparable, if not somewhat lower than the majority of mammalian NOAELs. Acute toxicity endpoints for avian species exposed to Pb, via oral gavage show mortality ranging from around 125 mg/kg/d for a 35day study on chickens (Vengris and Mare´, 1974) to 625 mg/kg/d for American kestrels (Hoffman et al., 1985). Growth endpoints for avian species, are more similar to the NOAELs generated for fence lizards (31.5 mg/kg/d) with a range of NOAELs for growth from 2.5 to 61.3 mg/kg/d for a variety of species and study durations. There are few actual avian studies where exposure route and duration were similar to those used here. A study on American kestrels
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(Hoffman et al., 1985) where birds were exposed via gavage for 10 days showed a NOAEL for growth of 25.0 mg/kg/d. There were significant effects of Pb on all sub-lethal endpoints evaluated during the 60 d sub-chronic study including cricket consumption, body weight, hematological parameters, clinical chemistry parameters, histopathological observations and postdose behaviors for lizards. There were no strong age-related effects although effects related to body weight seemed greater for adults; for this parameter, growth associated with the juvenile stage likely impacted the response to Pb. While there are some acute and subacute toxicity data for reptiles exposed to Pb, we could not compare the data collected from the present 60 d study with the results of other studies because there are no data on the toxicity of Pb to reptiles for exposures longer than two weeks. In addition, there was a significant interaction between time and dose suggesting that the effects of Pb became more pronounced with duration of exposure; this bears out from visual inspection of the curves. Juveniles clearly grew through the course of the study so any effects on change in body weight are superimposed over changes in body weight associated with somatic growth and maturation. To address the well-recognized shortcomings of using NOAELs and LOAELs (Laskowski, 1995), we conducted a benchmark dose analysis (BMD; Filipsson et al., 2003) to determine the exposure concentration associated with a 10% effect level on body weight. A 10% effect level was chosen because it is thought to be near what might be considered an ecologically meaningful and measurable effect. For adults the exposure level estimated to result in a 10% reduction in body weight was 5.9 mg/kg/d while in juveniles the exposure level was 8.9 mg/kg/d. These values are near the exposure concentration of 10 mg/kg/d which resulted in statistically significant effects on body weight in juvenile and adult lizards. Importantly, the BMD analysis is based on a 60-day exposure although there is an apparent effect of time as well. For shorter duration exposures, it is possible that the BMD for a 10% effect level would be higher. Presently, the U.S. Environmental Protection Agency Ecological Soil Screen Levels (Eco-SSLs) for Pb are 11 mg/kg dry weight in soil for avian wildlife and 56 mg/kg dry weight in soil for mammals (US EPA, 2005a). The results for the BMD analysis for lizards approximate the avian Eco-SSL which, was based on risk to avian ground insectivores (e.g. woodcock). In addition to effects on body weight and food consumption, there were significant effects of Pb on organ weight, hematological parameters and post-dose behaviors. Examination of the organ weight data shows that there were significant effects on kidney weight with a positive relationship to Pb dose and fat and gonad weight, which showed a negative relationship to Pb dose. These results suggest that there may be effects of Pb on the kidney and also that energy acquisition or allocation may have been affected as manifested by a decrease in gonad and body fat weight. Classic signs of Pb induced anemia were observed with decreased hemoglobin at both 10 and 20 mg/kg/d. Hematocrit, total solids, calcium, and glucose were also reduced at 20 mg/kg/ d. Although, adults had decreased RBCs at 10 and 20 mg/kg/d and the juveniles did not, decreases in albumin and total protein were observed in juveniles at 20 mg/kg/d and not in the adults. No effects on hematology or clinical chemistry were observed in the 1.0 mg/kg/d treatment. Post-dose observations indicated that there were dose-related effects on gaping, arched posture and dark coloration. Although the ecological significance of these behaviors is unknown, they warrant further study as alterations in behavior in wild animals may have significant survival consequences. Taken as a whole, there was little observed mortality; however, the sub-lethal effects observed are likely to result in negative effects on survival and reproduction were similar effects to manifest in the field.
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There is a well-recognized lack of data on the toxicity of chemicals to reptilian species (Hopkins, 2000). Research efforts have, for the most part, focused on measuring body burdens in field-collected reptiles, primarily chelonians (Sparling et al., 2000). Importantly, while the toxicity testing system using S. occidentalis seems highly suitable for routine use in developing reptile toxicity data as exemplified by this study and other recently published work (Holem et al., 2006; Suski et al., 2008; McFarland et al., 2008), the applicability of these toxicity data towards other reptilian species remains unknown. Since turtles, snakes and other lizard species can have very different life histories, behaviors, and habitat preferences compared to S. occidentalis, any data generated using this species must be applied with caution when evaluating risks. Nonetheless, recent efforts in reptile toxicity testing (Suski et al., 2008; McFarland et al., 2008) marks an advance in furthering our understanding of the effects of xenobiotics on reptiles by providing a robust, simple, and efficient testing system. Although more reptile toxicity data are becoming available, significant questions remain regarding exposure to reptiles under natural conditions. Future studies regarding the potential risks of contaminants to reptiles should address this source of uncertainty. Exposure studies and estimation methods should reflect the biology of this class of animals with explicit consideration of dermal exposure (many reptiles are in close proximity to the ground), incidental or intentional soil ingestion (Beyer et al., 1994a,b; Rich and Talent, 2009) and bioavailability from soil and food items. Moreover, because reptiles are ectothermic, energy demands and hence consumption rates will be environmentally dependent and exposure estimation methods must account for this fundamental physiological difference compared to birds and mammals. Acknowledgements The authors would like to thank anonymous reviewers for helpful comments. All research was conducted at U.S. Army Center for Health Promotion and Preventive Medicine. This study was conducted consistent with Good Laboratory Practices and through and approved protocol with the Institutional Animal Care and Use Committee. The animal exposures and husbandry were performed in animal facilities fully accredited by the American Association for the Assessment and Accreditation of Laboratory Animal Care. The views expressed in this paper are those of the authors and do no necessarily reflect the views and policies of the U.S. Army. References Beyer, W.N., Connor, E., Gerould, S., 1994a. Survey of soil ingestion by wildlife. The Journal of Wildlife Management 58, 375–382. Beyer, W.N., Connor, E.E., Gerould, S., 1994b. Estimates of soil ingestion by wildlife. The Journal of Wildlife Management 58, 375–382. Brasfield, S.M., Weber, L.P., Talent, L.G., Janz, D.M., 2002. Dose-response and time course relationships for vitellogenin induction in male Western fence lizards (Sceloporus occidentalis) exposed to ethinylestradiol. Environmental Toxicology and Chemistry 21, 1410–1416. Campbell, K.R., Campbell, T.S., 2002. A logical starting point for developing priorities for lizard and snake ecotoxicology: a review of available data. Environmental Toxicology and Chemistry 21, 894–898. Filipsson, A.F., Sand, S., Nilsson, J., Victorin, K., 2003. The benchmark dose methodreview of available models, and recommendations for application in health risk assessment. Critical Reviews in Toxicology 33, 505–542.
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