Ecological risk of anthropogenic pollutants to reptiles: Evaluating assumptions of sensitivity and...

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Environmental Pollution xxx (2010) 1e11

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Environmental Pollution

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Ecological risk of anthropogenic pollutants to reptiles: Evaluatingassumptions of sensitivity and exposure

Scott M. Weir a, Jamie G. Suski b, Christopher J. Salice a,*

a Texas Tech University, The Institute of Environmental and Human Health, Department of Environmental Toxicology, Box 41163, Lubbock, TX, USAb Texas Tech University, Department of Biological Sciences, Box 43131, Lubbock, TX, USA

Avian receptors are not universally appropriate surrogates for reptiles

a r t i c l e i n f o

Article history:Received 2 April 2010Received in revised form29 July 2010Accepted 16 August 2010

Keywords:ReptileEcological risk assessmentExposureContaminantsDermal

* Corresponding author.E-mail addresses: scott.weir@ttu.edu (S.M. Weir), ja

chris.salice@ttu.edu (C.J. Salice).

0269-7491/$ e see front matter � 2010 Elsevier Ltd.doi:10.1016/j.envpol.2010.08.011

Please cite this article in press as: Weir, S.M.and exposure, Environmental Pollution (201

a b s t r a c t

A large data gap for reptile ecotoxicology still persists; therefore, ecological risk assessments of reptilesusually incorporate the use of surrogate species. This necessitates that (1) the surrogate is at least assensitive as the target taxon and/or (2) exposures to the surrogate are greater than that of the targettaxon. We evaluated these assumptions for the use of birds as surrogates for reptiles. Based on a survey ofthe literature, birds were more sensitive than reptiles in less than 1/4 of the chemicals investigated.Dietary and dermal exposure modeling indicated that exposure to reptiles was relatively high, particu-larly when the dermal route was considered. We conclude that caution is warranted in the use of avianreceptors as surrogates for reptiles in ecological risk assessment and emphasize the need to betterunderstand the magnitude and mechanism of contaminant exposure in reptiles to improve exposure andrisk estimation.

1. Introduction

Reptiles are considered a globally declining taxon (Gibbonset al., 2000). Hypotheses for reptile decline include: habitat lossand degradation, invasive species, disease/parasitism, globalclimate change, and environmental pollution (Gibbons et al., 2000).Our understanding of pollutant effects on reptiles is still lackingdespite long-standing and well-recognized data gaps (Hopkins,2000; Sparling et al., 2010). In general, past reptile studies havefocused on measuring body burdens of various pollutants fromsamples collected in the field (Meyers-Schöne, 2000). While thesedata are useful for understanding historical exposures of givenpopulations, the actual risks and population-level effects ofpollutants on reptiles is still largely unknown and generallyunderstudied. Aside from research on the effects of endocrinedisrupting compounds on alligators and snapping turtles (reviewedin Willingham, 2005), some progress has been made regardingcertain contaminant groups such as explosives (McFarland et al.,2008, 2009; Suski et al., 2008), organophosphates (Holem et al.,2006, 2008; Hall and Clark, 1982; Sanchez et al., 1997; Sánchez-Hernández and Walker, 2000; Bain et al., 2004), and heavy

mie.suski@ttu.edu (J.G. Suski),

, et al., Ecological risk of anthr0), doi:10.1016/j.envpol.2010

metals (Holem et al., 2006; Burger et al., 1998; Hopkins et al., 1999,2002, 2005; Marco et al., 2004; Rich and Talent, 2009; Jones andHolladay, 2006; Brasfield et al., 2004; Mann et al., 2007; Saliceet al., 2009). Also, efforts to develop and utilize a reptile toxicity-testing model (Talent et al., 2002; McFarland et al., 2008, 2009;Suski et al., 2008; Salice et al., 2009) have been largely successfulalthough no large-scale data development has ensued.

The abovementioned studies focus largely on understandingand/or developing methods to understand the effects of pollutantson reptiles. These data are essential for risk assessments butrepresent only half of the equation, as risk is a function of both theeffects of a given pollutant and the exposure to that pollutant. Areview of the literature on reptile ecotoxicology indicates that fewstudies have focused on understanding the magnitude and mech-anism of contaminant exposure in reptiles (Hopkins et al., 2002).Generally, in ecological risk assessment (ERA), the most commonlyconsidered route of exposure is dietary. Despite multiple reports ofsignificant dermal exposure in birds (Vyas et al., 2007; Driver et al.,1991) the dermal route is less likely to be considered for ecologicalrisk assessment, owing to a general lack of dermal toxicity orexposure data. Dermal toxicity studies are not required forecological receptors under the Federal Insecticide Fungicide andRodenticide Act (FIFRA). If the dermal route is explicitly considered,it is possible that reptiles would receive higher relative exposurethan birds, due to greater percentage body contact with soil andvegetation. Direct assessment of the relative contribution of dermal

opogenic pollutants to reptiles: Evaluating assumptions of sensitivity.08.011

S.M. Weir et al. / Environmental Pollution xxx (2010) 1e112

exposure to total exposure of reptiles has not been reported,however a few studies indicate that dermal exposure could besignificant (Talent, 2005; Brooks et al., 1998a,b; Alexander et al.,2002).

As is common for most receptors in ERAs, avian exposure isestimated using mathematical models. For example, the methodused by the United States Environmental Protection Agency(USEPA) Office of Pesticide Programs for Tier 1 risk assessmentsincorporates avian allometric relationships for food consumptionrates (USEPA, 1993) to estimate dietary exposure (USEPA, 2004;Terrestrial Residue ExposureModel [T-REX]). Generally, dermal andinhalation exposures are not explicitly considered relative to thedietary exposure pathway except under certain scenarios (Sampleet al., 1997). Alternatively, because conservative assumptions areused in estimating dietary exposures (e.g., 100% contaminated diet),dermal and inhalation exposure can be ignored. In this case, high-end estimation of dietary exposure would be considered protectiveof dermal and/or inhalation exposure. Whether conservativeassumptions related to dietary exposure are protective is relativelyunderstudied although limited data suggest that dermal exposurein birds may be significant (Driver et al., 1991; Henderson et al.,1994: Vyas et al., 2007). An example is the past use of pesticides/deterrents to coat perching areas to control avian pests (Clark,1997;e.g. “rid-a-bird” perches, Jackson, 1978).

As with birds, contaminant exposure to reptiles could occur viadietary, drinking water, dermal, and/or inhalation routes. Soil inges-tion has been shown to be a possible source of contaminants ingeckos (Rich and Talent, 2009), whichmay also be applicable to othersquamates (and wildlife in general, Beyer et al., 1994) and generallyfalls under dietary exposure. Dermal exposure of reptiles has beenstudied sparingly (Talent, 2005; Brooks et al., 1998a,b), makingchemical-specific toxicity comparisons to other routes of uptake andmechanistic understanding difficult. Dermal exposure may besignificant for reptiles that maintain considerably greater proximityto soil than birds, although a common view is that the permeabilityand thickness of reptile epidermis, which can contain significantamounts of keratin and lipids (Pough et al., 2004), may reduce oreliminate exposure via the dermal route (Snodgrass et al., 2008).However, skin permeability is more likely affected by lipid content,not keratin, which offers greater resistance to water movement(Roberts and Lillywhite,1980; Palmer, 2000). If a high lipid content oflizard epidermal tissue prevents water loss, it should consequentlyallow increased absorption of lipophilic material. Conversely, a lowlipid content of reptile epidermismay be permeable to water solublecontaminants. The permeability of reptile epidermis to variouscontaminants has not been well studied (Hopkins, 2006).

The lack of data regarding exposure and effects of pollutants onreptiles presents a challenge for assessing the ecological risks posedby pollutants to reptile populations. In the face of significantuncertainties and data gaps, risk assessors frequently use a numberof methods for deriving risk estimates. Among these, the use ofsurrogate species is common (USEPA, 2004). As an example, birdshave been used as surrogates for reptiles in ERA, both currently andhistorically (Sample et al., 1997) under the assumption that expo-sure to birds is likely higher than reptiles; therefore, avian riskestimates are protective of reptiles. The basis of this assumption isthat because reptiles have lower energy demands than birds, die-tary exposures to reptiles are likely to be less because of a lowerfood intake compared to birds. Although these assumptions areevident in some ERAs and perhaps in the field at large, they havenot been fully reviewed or tested. In order for birds to be consid-ered suitable surrogates for reptiles in ERA the following criticalassumptions must be met: (1) that birds are equally or moresensitive than reptiles to contaminants and/or (2) that contaminantexposure is higher because of greater avian energetic demands.

Please cite this article in press as: Weir, S.M., et al., Ecological risk of anthrand exposure, Environmental Pollution (2010), doi:10.1016/j.envpol.2010

We review these two critical assumptions regarding the use ofavian receptors as surrogates for reptilian receptors in ERAs. Toevaluate whether birds are more sensitive than reptiles (assump-tion one), we compared acute toxicity estimates for the samecompound to both receptors. To evaluate whether exposure ishigher in birds than reptiles (assumption two), we used dietary anddermal exposure models to determine under what circumstancesand/or assumptions might exposure be unequal between birds andreptiles. Finally we conclude with a discussion of the relativeimportance of different exposure routes to risk estimation, andprovide recommendations with regard to future research efforts inreptile ecotoxicology.

2. Methods

2.1. Comparison of avian and reptile contaminant sensitivity

A literature search was conducted for contaminant toxicity data for chemicalscommon to both avian and reptilian receptors. Laboratory toxicity tests thatreported mortality endpoints were used for comparison to reduce ambiguitiesassociated with field studies, and behavioral or reproductive endpoints in differenttaxa. Avian and reptilian studies were used for comparison if there was reasonableagreement in route of exposure and study duration. In some cases, the paucity ofreports on reptiles required comparing two studies with dissimilar routes ofexposure or duration period to generate a sense of the relative taxa-specificsensitivity.

For reports dated 1997 and earlier, two sources for toxicity values were used, theReptile and Amphibian Toxicology Literature database (Pauli et al., 2000) and forlizards specifically, Campbell and Campbell (2000). For more recent reports onreptile toxicology, Google Scholar and the Science Citation Index were searchedextensively. It was quickly discovered that restricting search phrases would severelylimit the number of reptile studies obtained. Many studies were found fortuitouslyusing multiple Boolean modifiers, for example: “(toxic* OR effect*) AND (reptile* ORlizard* OR snake* OR turtle*).”

Avian toxicity review papers and databases were used as a source of informationfor multiple species and chemicals. Schafer et al. (1983), Tucker and Haegele (1971),andMineau et al. (2001) provide a number of toxicity values for birds. In cases whereno (or few) published data were found, we used the Pesticide Ecotoxicity Database(PED, 2010). Reports included in the PED are peer-reviewed but not be necessarilyavailable in the primary literature (e.g., data required for pesticide registration underFIFRA).

2.2. Comparison of avian and reptile exposure to contaminants

Exposure models based on avian and reptile physiology and ecology were usedto evaluate the relative expected dietary exposures to birds and reptiles. In addition,because there is no comprehensive understanding of how dermal exposurecontributes to toxicity in reptiles and because available evidence suggests it may beimportant (Brooks et al., 1998a,b), we constructed a dermal exposure model tobetter understand the potential impact of dermal exposure to total exposure inreptiles.

Our dietary exposure models were based on estimates of daily energy require-ments obtained from animals in the field (Sample et al., 1997; USEPA, 1993, 2004).Specifically, allometric equations of fieldmetabolic rate (FMR, Nagy et al., 1999) wereused to generate body-weight (BW) specific energy demands, and then subsequentdoses (via ingestion) and risk estimates. We used two different equations for eachtaxa to generate a range of potential exposures. The allometric equations used forreptiles were:

FMR ðkcal=dÞ ¼ 0:0468�BWðgÞ0:889

�(1)

�(2)

Equation (1) was calculated as an average for all reptiles, and equation (2) representsa reptile with a larger metabolic demand (i.e., lizards in the family Lacertidae).

For birds, the following allometric equations were used:

�(3)

�(4)

In which equation (3) is an average metabolic rate for all birds, and equation (4)is the average for birds in the order Galliformes (Nagy et al., 1999). The order Gal-liformes includes the bobwhite quail, used in toxicity testing and registration ofpesticides under FIFRA.

S.M. Weir et al. / Environmental Pollution xxx (2010) 1e11 3

The resulting equation for estimating dietary dose (DDiet) was:

DDietðmg=kg BW=dÞ ¼ FMRðkcal=dÞ=ðBWðgÞðCinsectsðmg=kgÞÞ=ð1:7 kcal=gÞ (5)

whereCinsects was the concentration of contaminant residues on insects.We assumedthat all species ingested insects (1.7 kcal/g) and insect residues were 45 mg/g(corresponds to a 1 lb/acre pesticide application rate based on the Kenaganomogram,T-REX).

We developed a dermal exposure model for birds and lizards that was based onan adapted version of the methods for exposure estimation used by USEPA (T-REXand EPA’s Probabilistic Risk Assessment model [TIM]: http://www.epa.gov/opp00001/science/models_pg.htm) and is similar to a previously published model(Hope, 1995). Our dermal exposure model was:

DDermalðmg=kg BW=dÞ ¼�Kpðcm=hÞ

�0:30 SA

�cm2�=HðcmÞ

� Tðh=dÞCSoil�mg=m2�BF

�.BWðgÞ ð6Þ

where DDermal is the calculated dose associated with dermal exposure. Kp is thechemical-specific permeability coefficient estimated from Walker et al. (2003). Thevalue of Kp we selected represents an average value for all chemicals that were moretoxic to reptiles than to birds (see results: Literature survey: effects). SA stands forsurface area; for lizards it wascalculated using a relationship derived for salamanderas a proxy SA¼ 8.42*BW0.694 with 30% of the total used for dermal exposure, forbirds SA¼ 10*BW0.667 with 7% of the total used to represent dermal exposurethrough feet (USEPA, 1993). T is the duration of contact between the organism andcontaminated soil and is assumed to be 12 h per day. This value does not representthe total time in contact with soil, which could be more or less depending onpreferred habitat (e.g. arboreal lizards and ground-dwelling snakes). CSoil is the soilconcentration based on 1 lb/acre application spread evenly over 1 acre of soil(112 mg/m2) and mixed evenly to a depth of 1 cm. It was assumed that only the firstmm of soil was in contact with skin, therefore Csoil¼ 11.2 mg/m2. BF is an absorptionfactor adjustment to account for the bioavailability of contaminants from soil. Skinthickness, H, was estimated as1% of the body radius (r) calculated as:

r ¼�� hþ sqrt

�h2 þ ð2pSAÞ

��.ð2pÞ (7)

in which h represents the height of a cylinder which is equivalent to the snout-ventlength of a lizard. The value of 1% was chosen so that the snout-vent length datareported by Pough (1973), for multiple species of lizards, would provide results thatwere similar to reported skin thickness in geckos (i.e. 0.22e0.36 mm, Bauer et al.,1989), A range of skin thickness was created using snout-vent lengths that corre-sponded to body weights used in our models. No skin thickness estimates for avianfeet were found; therefore, this value was set as the mean of skin thicknesses usedfor reptiles: 0.5 mm.

To better understand the ecological risk implications of differential toxicity andexposure in avian and reptilian receptors, we generated estimates of risk in the formof hazard (or risk) quotients. These are standard, unit less values obtained fromdividing exposure estimates by toxicity estimates and provide a screening-levelassessment of the potential for adverse effects. In our simulations we used chemicalsidentified in the literature survey as being more toxic to reptiles than birds (rote-none, pyrethrins, and diphacinone) since, under these circumstances, birds may beused as surrogates for reptiles and hence, should be protective. For chemicals withthe greatest difference in acute toxicity between reptiles and birds, we determinedan average relative sensitivity for reptiles compared to birds by taking the mean ofthe ratio between reptile and bird Lethal Doses (LD50s) (Table 1). For example, theavian LD50 for pyrethrins is 5620 mg/kg and for reptiles it is 15 mg/kg, the ratio ofthese two values is 374.6. This was done for pyrethrins, rotenone, and diphacinone.The average was then used as an example of a “worst case scenario” in which birdswere used as surrogates for a theoretical chemical to which reptiles were 179timesmore sensitive (Table 1). Note that the intent of this modeling exercise was tocompare the exposure for avian and reptilian receptors and hence, while we usedreasonable parameter estimates and assumptions (e.g. residue concentrations and

Table 1Ratios of lethal endpoints for reptiles and birds used in risk and exposure models.The lowest avian endpoints were used when multiple LD50 values were available.Ratios represent a measure of relative risk to reptiles compared to birds. Forexample, pyrethrins are 374 times more toxic to reptiles than birds. The average ofthe ratios was then used to represent a hypothetical chemical to which reptiles were179 times more sensitive than birds.

Chemical Lowestavian LD50

ReptileLD50

Ratio (Bird:ReptileLD50)

Pyrethrins 5620 15 374.6Rotenone 113 2 56.5Diphacinone 3158 30 105.3Hypothetical Chemical e e 178.8

energy content of insects), we recognize variation or more detailed informationregarding parameters can have significant impacts on model results andconclusions.

3. Results

3.1. Literature survey: effects

A survey of the literature yielded 17 reports of laboratorytoxicity data involving reptiles exposed to 15 different chemicals.Thirteen reports containing data on the same 15 chemicals werelocated for birds. The results for birds and reptiles are summarizedin Table 2. Reptiles weremore sensitive to 5 out of 15 chemicals. For3 out of 15 chemicals, birds were more sensitive than reptiles. Forthe remaining 7 chemicals, the data found were either inconclusive(n¼ 2), or birds and reptiles were more or less equivalent in theirsensitivity to the chemical (n¼ 5). Inconclusive results wereattributable to differences in study design, exposure route, durationof study, or endpoints measured.

3.2. Explosives

An appreciable amount of data on the toxicity of military-relatedcompounds to both birds and reptiles has been generated. Lizardswere more sensitive than birds to acute doses of RDX (McFarlandet al., 2009; Gogal et al., 2003), and subchronic doses of TNT(McFarland et al., 2008; Johnson et al., 2005a; Gogal et al., 2002).For both acute and subchronic doses of DNT, birds were moresensitive than reptiles (Suski et al., 2008; Johnson et al., 2005b).

3.3. Cholinesterase inhibitors

Toxicity data from four organophosphate (OP) insecticides(malathion, parathion, propoxur, and fenitrothion) suggested thatbirds were either more sensitive to OPs or were of equivalent sensi-tivity to reptiles. Birds appeared to be much more sensitive to para-thion than reptiles (Tucker and Haegele, 1971; Mineau et al., 2001;Schafer et al., 1983; Hall and Clark, 1982; Yawetz et al., 1983). Feni-trothion results were variable, and there was little reptile data onwhich to make comparisons. Data on cholinesterarse inhibition byfenitrothion showed that the zebra finch (Poephila guttata) wasmoresensitive (Holmes and Boag, 1990) than the central bearded dragon(Pogona vitticeps) (Bain et al., 2004). For mortality associated with-fenitrothion exposure, Bain et al. (2004) dosed central beardeddragons with up to 20mg/kg without any mortality, while themedian LD50 of 12 avian species was 63.43 mg/kg (Mineau et al.,2001). Additionally, the LD50s of red-winged blackbird (Agelaiusphoeniceus), European starlings (Sturnus vulgaris), and coturnix quail(Coturnix coturnix) were 20.5, 11, and 56mg/kg, respectively, sug-gesting that birds were more sensitive to fenitrothion than reptiles(Schafer et al., 1983). Birds and reptiles appeared to have equivalentsensitivities to malathion (Mineau et al., 2001; Schafer et al., 1983;McEwen and Brown, 1966; Hall and Clark, 1982; Özelmas and Akay,1995; Holem et al., 2006, 2008) and propoxur (Schafer et al., 1983;Tucker and Haegele, 1971; Mineau et al., 2001; Brooks et al., 1998a).

3.4. Other contaminant groups

Pyrethrins (and their pyrethroid derivatives) are a group ofneurotoxic insecticides that appeared much more toxic to reptilesthan birds (Talent, 2005; Brooks et al., 1998a,b; PED, 2010).Reportsof avian toxicity from pyrethroids were uncommon due the generallow to moderate toxicity of this class of chemicals to avian recep-tors. This is reflected, for example, in the high avian LD50of>5000 mg/kg for pyrethrins (PED, 2010).

Table 2Summary of comparable reptile and bird acute toxicity literature.

Chemical Conclusionsa Speciesb Exposure Duration Endpoint Results Referencee

RDX R Bobwhite Oral Acute (14 day observation) LD50 187 mg/kg (males) 280 mg/kg (females) 1Western Fence Lizard Oral Acute (14 day observation) LD50 72 mg/kg (males) 88 mg/kg (females) 2Bobwhite Dosed Feed Subchronic (90 days) Survival, body weight,

food consumptionFeed consumption and weight lossdecreased initially, but birds acclimated bythe third week. No significant decreasein survival

Western Fence Lizard Oral Subchronic (60 days) Survival, growth rate,food consumption

�5 mg/kg/d¼ decreased growth and foodconsumption �8 mg/kg/d¼ decreased survival

TNT R Pigeon Oral 60 day Chromaturia prevalence,Body Weight, and FeedConsumption

�70 mg/kg¼ increased chromaturia,200 mg/kg¼ decreased body weight, nodecrease in food consumption

Western Fence Lizard Oral 60 day Chromaturia prevalence,Body Weight, and FeedConsumption

�35 mg/kg¼ increased chromaturia,45 mg/kg¼ decreased body weight,�35 mg/kg¼ decreased feed consumption

Bobwhite Dosed Feed 90 day Body weight, Feedconsumption,Hematology

No changes in body weight or feedconsumption, blood parameters decreasedin a dose-response pattern

DNT B Bobwhite Oral Acute LD50 55 mg/kg 6Western Fence Lizard Oral Acute LD50 380 mg/kg (males) 577 mg/kg (females) 7Bobwhite Oral Subacute (14 day) Mortality �35 mg/kg¼ 100% Mortality 6Western Fence Lizard Oral Subacute (14 day) Mortality 33% Mortality in 100 mg/kg/d 100%

Mortality in 200 mg/kg/d7

Malathion E Sharp-tailed Grouse Oral 25e60 days Mortality 16.7% mortality in 171-200 and 221-240.50% mortality in 221-240.

Redwinged Blackbird Oral Acute (duration not reported) LD50 LD50¼ 400 mg/kg 98 avian species (unspecified) Oral Acute (duration not reported) Median LD50 Median LD50¼ 466.5 mg/kg 10Dwarf Lizard Not Reported Acute (duration not reported) LD50 LD50¼ 169.8 mg/kg 11Green Anole Oral 24 hours 24 hr LD50 LD50¼ 2324 mg/kg 12Western Fence Lizard Oral Sing dose, 13 days observation Mortality 20% mortality in the 200 mg/kg dose group 13Western Fence Lizard Oral Dose every 27 days (81 day

observation)Mortality 8% mortality in the 20 mg/kg dose group,

23% mortality in the 100 mg/kg dose group14

Parathion E 6 avian species Oral Observed for 14 days LD50 Mean LD50¼ 8.39 mg/kg min¼ 2.13max¼ 24.0

19 avian species (unspecified) Oral Acute (duration not reported) Median LD50 Median LD50¼ 5.62 mg/kg 107 avian species Oral Acute (duration not reported) LD50 Mean LD50¼ 3.18 mg/kg min¼ 1.33

max¼ 5.629

Green Anole Oral 24 hours LD50 LD50¼ 8.9 mg/kg 12Caspian Terrapin Oral 4 Days LD50 LD50¼ 10 mg/kg 16

Fenitrothion B Zebra Finch Oral Acute (10 day observation) Plasma ChE inhibition,mortality

All doses caused significant plasma ChEinhibition up to 12 hrs post dose. 10 and30% mortality in 4 and 11 mg/kg dosegroup, respectively

12 avian species (unspecified) Oral Acute (duration not reported) Median LD50 Median LD50¼ 63.43 mg/kg 105 avian species Oral Acute (duration not reported) LD50 Mean LD50¼ 144 mg/kg min¼ 11

max¼ 3169

Central Bearded Dragon Oral Acute (3 week observation) Plasma ChE inhibition High dose group (20 mg/kg) had significantlyinhibited plasma ChE for up 5 dayspost-dose. No mortality.

Propoxur E 6 avian species Oral Observed for 14 days LD50 Mean LD50¼ 26.2 mg/kg min¼ 11.9max¼ 60.4

10 avian species Oral Acute (duration not reported) LD50 Mean LD50¼ 14.2 mg/kg min¼ 3.83max¼ 42.2

23 avian species (unspecified) Oral Acute (duration not reported) Median LD50 Median LD50¼ 11.76 mg/kg 10Brown Tree Snake Oral 7 days observation Mortality LD50¼ 15 mg/kg (estimated from data) 20

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Pyrethrins R Bobwhite Dosed Feed 8 days LC50 LC50¼ 5620 mg/kg 19Green Anoleb,c Dermal 2 second dermal exposure, 48 hour

observationLC50 LC50¼ 77.6 mg/L (@ 20 C) 45% mortality for

300 mg/L (@ 35 C)21

Brown Tree Snake Oral 72 Hours observation Mortality LD50¼ 15 mg/kg (estimated from data) 20

Fipronil E Zebra Finchc,d Oral Acute (4 week observation) LD50 LD50¼ 45.41 mg/kg 22Ring-necked Pheasant Oral 14 days LD50 LD50¼ 31 mg/kg 197 avian species (unspecified) Oral Acute (duration not reported) Median LD50 Median LD50¼ 39.19 mg/kg 10Bobwhite Oral 21 days LD50 LD50¼ 1065 mg/kg 19Fringe-toed lizardc,d Dosed feed Acute (4 week observation) Mortality 50% mortality at 30 mg/kg 23

Diphacinone R Mallard Oral 14 days observation LD50 LD50¼ 3158 19Brown Tree Snake Oral 7 days observation Mortality LD50¼ 30 mg/kg (estimated from data) 20

Sodium Fluoroacetate(1080)

B 6 avian species Oral Observed for 14 days LD50 Mean LD50¼ 7.33 mg/kg min¼ 3.0max¼ 17.7

2 Species (Red-winged blackbird,European Starling)

Oral Acute (duration notreported)

LD50 Blackbird LD50¼ 4.22 mg/kg StarlingLD50¼ 2.37 mg/kg

55 avian species (unspecified) Oral Acute (duration not reported) Median LD50 Median LD50¼ 5.46 mg/kg 10Shingleback skink Intraperitoneal

Injection15 day observation LD50 LD50 for naïve skinks¼ 200 mg/kg 24

Fluoro-acetamide(1081)

I Multiple birds of prey Consumption ofdosed prey

8 - 21 days Mortality Mortality in one heron after 21 daysof a 3.4 mg/kg dose, otherwise, nomortality to any species after0.2-3.4 mg/kg dose

2 avian species (unspecified) Oral Acute (duration notreported)

Median LD50 Median LD50¼ 9.46 mg/kg 10

Multiple snakes Consumption ofdosed prey

1 - 41 days Mortality No mortality in any snakes exposedto 0.1-0.8 mg/kg

Rotenone R Mallard, Ring-necked Pheasant Oral 14 days observation LD50 Mallard LD50¼>2200 mg/kgPheasant LD50¼ 1680

9 avian species Oral 12-24 hours LD50 LD50 ranged from 113 - 3000 mg/kg 26Brown Tree Snake Oral 72 Hours observation Mortality LD50¼ 2 mg/kg (estimated from data) 20

Methyl-Thiophanate I Bobwhite Oral 8 days LD50 LD50¼ 4640 mg/kg 19Lizard (Podarcissicula) Intraperitoneal

InjectionAcute (15 day observation) LD50 LD50¼ 900 mg/kg 27

Lead (Lead Acetate) E Bobwhite (6 week old) Dosed Feed 6 weeks Mortality 17, 19, and 47% mortality in 1000,2000, and 3000 ppm treatments,respectively

Western Fence Lizard Single oral dose 13 days Mortality 30% mortality in 1000 mg/kg dose group 13Western Fence Lizard Single oral dose < 7 days Approximate

lethal doseALD¼ 2000 mg/kg 29

Red-eared Slider (hatchlings) Single Oral dose 4 months Mortality LD50¼ 500 mg/kg NOEL¼ 100 mg/kg 30

a R¼ Reptilian receptor more sensitive, A¼Avian receptormore sensitive, E¼ Equivalent sensitivity, I¼ Inconclusive given the data.b Species listed: Bobwhite (Colinus virginianus), Western fence lizard (Sceloporus occidentalis), Pigeon (Columbia livia), Sharp-tailed grouse (Tympanuchus phasianellus), Red-winged blackbird (Agelaius phoeniceus), Dwarf lizard

(Lacerta parva), Green anole (Anolis carolinensis), Caspian terrapin (Mauremys caspicarivulata), Zebra finch (Poephila guttata), Central bearded dragon (Pogona vitticeps), Brown tree snake (Boiga irregularis), Ring-necked pheasant(Phasianus colchicus), Fringe-toed lizard (Acanthodactylus dumerili), Mallard (Anas platyrhynchos), European starling (Sturnus vulgaris), Shingleback skink (Tiliqua rugosa).

c Green anoles were exposed to a formulation of pyrethrins that included a pirperonylbutoxide synergist.d A formulation of fipronil was used in these studies, in which a dispersal solvent may add to toxicity.e References: 1. Gogal et al., 2003, 2.,McFarland et al., 2009, 3. Johnson et al., 2005b, 4. McFarland et al., 2008, 5. Gogal et al., 2002, 6. Johnson et al., 2005a, 7. Suski et al., 2008, 8. McEwen and Brown, 1966, 9. Schafer et al., 1983,

10. Mineau et al., 2001, 11. Özelmas and Akay, 1995, 12. Hall and Clark, 1982, 13. Holem et al., 2006, 14.Holem et al., 2008, 15. Tucker and Haegele, 1971, 16. Yawetz et al., 1983, 17. Holmes and Boag, 1990, 18. Bain et al., 2004, 19.PED, 2010, 20. Brooks et al., 1998a, 21. Talent, 2005, 22. Kitulagodage et al., 2008, 23. Peveling and Demba, 2003, 24. Twigg and Mead, 1990, 25. Braverman, 1979, 26. Cutkomp, 1943, 27. Sciarrillo et al., 2008, 28. Damron andWilson, 1975, 29. Salice et al., 2009, 30. Burger et al., 1998.

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0 100 200 300 400

001051

Body Weight (g)

)d/lack(sdna

ygrenE

All BirdsGalliformesAll ReptilesLacertid Lizards

Fig. 1. Energy demands (kcal/d) for birds and reptiles as a function of body weight (g)based on allometric equations (Nagy et al., 1999). See text for models used for birds andreptiles. “All birds” and “All reptiles” represent average equations from multiplespecies across multiple taxa for each group. “Galliformes” and “Lacertid lizards” referto average equations for species from the specific group represented. Galliformesrepresent the bird group with the least metabolic demand, while lacertid lizardsrepresent the group of reptiles with the greatest metabolic demand.

0 100 200 300 400

Body weight (g)

)d/gk/gm(

Fig. 2. Estimated daily dietary dose (mg/kg/d) of a contaminant for birds and reptilesbased upon allometric equations used in Fig. 1 and prey common to terrestrial birdsand reptiles (insect, 1.7 kcal/g) with a body residue of 45 mg/kg. See text for completeexplanation of models.

Toxicity data for rotenone (a piscicide) indicated that reptilesmay have much greater sensitivity than birds. The estimated LD50for brown tree snakes exposed to rotenone was 2 mg/kg (Brookset al., 1998a), while LD50s from multiple avian species rangedbetween 113 and 3000 mg/kg (Cutkomp, 1943). Mallard and ring-necked pheasant LD50s were >2200 mg/kg and 1680 mg/kg,respectively (PED, 2010). For two additional chemicals, results aresparse and more data are needed for both birds and reptiles. First,diphacinone (anti-coagulant) appeared to be far more toxic toreptiles than birds (Brooks et al., 1998a; PED, 2010). Lead (in theform of lead acetate) also appeared to be equivalently toxic toreptiles and birds (Burger et al., 1998; Holem et al., 2006; Saliceet al., 2009; Damron and Wilson, 1975) but avian toxicity data forlead acetate is not common, with a greater historical focus on leadshot. Also, route of exposure differed between the lead studies,with oral gavage used in the reptile studies, and a feeding studyused for quail (Damron and Wilson, 1975).

For the remaining chemicals (fipronil, fluoroacetamide, sodiumfluoroacetate and thiophanate-methyl) birds or reptiles either hadequivalent sensitivity (fipronil), birds were more sensitive (sodiumfluoroacetate), or the interpretation of results was confounded bydifferences in study design (fluoroacetamide, thiophanate-methyl).Fipronil caused 50% mortality to fringe-toed lizards at a concen-tration of 30 mg/kg dosed in food (Peveling and Demba, 2003). Themedian LD50 for several avian species ranged from 31e45.41 mg/kg(Mineau et al., 2001; Kitulagodage et al., 2008; PED, 2010). The LD50reported for bobwhite quail was much higher than for other avianspecies (1065 mg/kg; PED, 2010). The LD50 of the fungicide thio-phanate-methyl, administered via intraperitoneal injection, to thelizard Podarcis sicula was 900 mg/kg (Sciarrillo et al., 2008), and4640 mg/kg for bobwhite quail (oral exposure, PED, 2010).However, route of exposure and duration (8 days for bobwhite, 15days for lizard) differed. Finally, one study of the toxic effects offluoroacetamide (1081, a rodenticide) on multiple avian predatorspecies and several snakes was conducted by exposing both groupsto dosed prey (Braverman, 1979). Concentrations were low(snakes< 0.8 mg/kg, birds< 3.4 mg/kg) and no mortality wasfound in any group, with the exception of one heron that died afterexposure to 3.4 mg/kg in food. The median oral LD50 of fluo-roacetamide to two unspecified species of birds was 9.46 mg/kg(Mineau et al., 2001). The results for 1081 could be expected to besimilar as those for sodium fluoroacetate (1080, also a rodenticide)in which birds appeared to be much more sensitive than reptiles(Tucker and Haegele, 1971; Twigg and Mead, 1990; Mineau et al.,2001).

3.5. Simulation modeling: exposure

Allometric equations of field metabolic rate (Nagy et al., 1999)showed that birds havemuch greater energy demands than reptiles(Fig. 1). Hence, if only dietary exposure is considered birds are likelyto receive a greater dose of a contaminant than reptiles from similardiets (e.g., 3e5 times greater, Fig. 2). This held true as body massincreased.

Estimates of risk in the form of hazard quotients providedadditional insight into the relative differences between birds andreptiles. For these simulations, we chose to average the proportionof bird:reptile LD50 data to use as a general placeholder fora chemical that is more toxic to reptiles than birds. We used thisapproach because suitable reptile data was available for only a fewchemicals (rotenone, pyrethrins, and diphacinone). Despite greateravian energy demands, risk estimates for dietary exposure werehigher for reptiles than birds owing to the greater toxicity to reptiles(Fig. 3). The results were consistent throughout a range of bodyweights for both taxa.

Dermal exposure modeling suggested that under the specifiedassumptions (e.g., surface area contact, same duration of exposure),reptiles have the potential to receive higher dermal exposures thanbirds (Fig. 4). Intuitively, as mass increases, the ratio of surface areato volume decreases and dermal exposure would likewise decrease(Fig. 4). Even if the fraction of contaminant that was bioavailablethrough dermal uptake was moderate (e.g., 10%) then dermaluptake would be much greater than dietary exposure for reptilesbelow 150 g and approximately equal to diet above 150 g (Fig. 4).

0 100 200 300 400

0.02.0

4.06.0

8.00.1

Body weight (g)

)05DL/eso

All BirdsGalliformesAll reptilesLacertid lizards

Fig. 3. Estimated risk for birds and reptiles from a contaminant. Risk is determined bydividing daily dose (Fig. 2) by the LD50 of the contaminant to reptiles or birds and isa Hazard (or Risk) Quotient. In this example, LD50data was determined by taking theaverage of the relative proportion of LD50s for birds and reptiles for rotenone, pyre-thrins, and diphacinone (see Table 1) to visualize the potential differences betweenreptiles and birds for chemicals which were specifically more toxic to reptiles thanbirds.

0 100 200 300 400

Body Weight (g)

)lioS ni elbaliavaoiB %( FB

Dermal = DietDermal = 1/2*DietDermal = 1/4*DietDermal = 1/8*Diet

002003

004005

)lioS ni elbaliavaoiB %( FB

Dermal = DietDermal = 1/2*DietDermal = 1/4*DietDermal = 1/8*Diet

For reptiles, the bioavailability factor that caused equal exposurevia dietary and dermal routes was between 2 and 10% usingequation (1) representing allometric data for all reptiles (Fig. 5a).Under the assumptions in the dermal model, dermal exposure inbirds was not a significant route of exposure relative to dietaryexposure (Fig. 5b) except when bioavailability was very high. This

0 100 200 300 400

Body Weight (g)

)d/gk/gm(

erusopxelamre

BF = 0.1BF = 0.05BF = 0.01Bird BF = 0.05All reptiles diet

Fig. 4. Estimated daily dermal dose (mg/kg/d) for a contaminant for reptiles with 3different bioavailability factors (BF) to account for uncertainty in the dermal exposuremodel (see text). A BF of 0.1 indicates that 10% of the contaminant in the soil wasavailable for uptake, and so on for 0.05 and 0.01. For comparison, lines representdermal exposures for birds and reptiles, while points indicated data for dietaryexposure for both reptile groups. A mean Kp was created by averaging the Kp of fourchemicals to which reptiles were more sensitive than birds: rotenone, pyrethrins, RDX,and TNT (see results: Literature Survey: Effects).

0 100 200 300 400

Body Weight (g)

Fig. 5. Relationship between the bioavailability factor (BF) and body weight fordifferent ratios of dermal to dietary exposure for reptiles (A) and birds (B). Thebioavailability factor represents the proportion of a contaminant in soil that isbioavailable for dermal uptake to a reptile, which addresses an aspect of uncertainty inmodeling dermal exposure in reptiles and birds. Note that as weight increased,a greater bioavailability was needed to maintain the ratio of dermal to diet, due todecreased surface area to volume ratio. Also note the much greater bioavailabilityneeded for birds due to a smaller surface area in contact with the soil (7%) compared toreptiles (30%).

was the result of a smaller surface area in contact with the soil (7%for birds compared to 30% for reptiles), and because the dietaryexposure was higher in birds than reptiles.

To gain a better sense of the conditions that determine therelative contribution of dietary and dermal exposure to totalexposure, we conducted a series of simulations in which we variedthe soil bioavailability factor (BF). As expected, as bioavailabilitydecreased, the contribution of dermal exposure to total exposuredecreased as well. If a moderate BF of 10% was assumed, whendermal and dietary exposures were combined the relationshipbetween bird and reptile exposure did not change greatly from thatof diet alone, although at low body weights, reptile exposure

showed a much greater increase relative to diet alone (Fig. 6a).However, if a contaminant is particularly toxic to reptiles but not tobirds, including dermal exposure can greatly increase risk toreptiles with almost no increase in risk to birds (Fig. 6b).

4. Discussion

Reptiles are among the least studied taxa in ecotoxicology. Thisis true despite the well recognized data gaps in understanding howenvironmental pollutants impact individuals and populations(Hopkins, 2000; Sparling et al., 2010) and concerns that environ-mental pollutants may contribute to reptile population declines(Gibbons et al., 2000). The lack of data on the effects of xenobioticsis likely due to lack of a regulatory requirement for toxicity testingon reptiles. The paucity of information on the mechanisms and

0 100 200 300 400

Body weight (g)

)d/gk/gm(

esoDla

0 100 200 300 400

Body weight (g)

)05DL/

erusopxelamre

Fig. 6. Exposure (A) and risk (B) estimates for dermal and dietary exposures combined.Risk was calculated as total exposure (dietþ dermal) divided by the LD50 for thecontaminant (LD50s are based on the average proportion of bird LD50 to reptile LD50,see Table 1). A moderate bioavailability factor of 10% was used to model dermalexposure for birds and reptiles. See text for models used.

relative magnitudes of xenobiotic exposure to reptiles may simi-larly be driven by a lack of legal requirements for data (e.g., FIFRA)and perhaps a greater appeal among researchers to evaluate effectsand/or body residues as opposed to exposure. As such, assessingrisks to surrogate organisms is one method for assessing risks toa taxon in the absence of data specific to that taxon. For reptiles,avian receptors have and can be used as surrogates for risk esti-mation based on the assumptions that (1) birds are as or moresensitive to contaminants than reptiles and (2) exposures arehigher in birds compared to reptiles because of higher avianenergetic demands. Our evaluations of these assumptions suggestthat, in light of current data (and lack thereof), these are not valid inall cases. Reptiles appear to be more sensitive than birds for somechemicals and consideration of dermal exposure may be importantfor both taxa but particularly for reptiles.

The lack of reptilian toxicity data limited our findings becausea large sample set of avian and reptilian chemical-specific toxicitydata were not available. The fact that for five out of fifteen chem-icals for which avian and reptile data were available, reptilesappeared to be more sensitive to contaminants than birds wassomewhat surprising and further corroborates the need for widertoxicity testing of reptiles. To that end, the use of theWestern fencelizard (Sceloporus occidentalis) as a toxicity-testing model may behelpful (Talent et al., 2002; Suski et al., 2008; McFarland et al.,2008; Salice et al., 2009).

In addition to differences in sensitivity to toxicants amongreptilian and avian receptors, the magnitude and nature of expo-sure may differ greatly between the groups, further confoundingefforts to use avian receptors as surrogates for reptilian risk. Thisresearch focus has received almost no attention with regards toreptiles, and furthermore, is generally lacking for birds as well.Specifically, we found that for dietary exposure only, birds did havehigher exposures, as expected, but that risks to reptiles couldexceed that of birds for compounds that are more toxic to reptiles.When we included dermal exposure, however, total exposure(dermalþ diet) to reptiles was similar to total exposure to galli-formes, but was still below total exposure levels for all birds(Fig. 6a). If a higher bioavailability was considered (e.g. 25e50%)then total exposure to reptiles could increase to levels equal toavian total exposure (data not shown). Importantly, the dermalexposure route could be significant for both reptiles (Talent, 2005;Brooks et al., 1998a,b; Alexander et al., 2002) and birds (Vyas et al.,2007; Driver et al., 1991; Henderson et al., 1994) and should receivegreater research attention.

As with any model, our dermal exposure models includedassumptions and highlighted uncertainties that may limit theapplication of suchmodels until further research is conducted. Firstwe assumed the surface area in contact with soil is the only route ofdermal exposure, and direct application of a pesticide or contactwith vegetation, was not included in the model, but could besignificant for both reptiles and birds (Driver et al., 1991; Brookset al., 1998a,b; Talent, 2005). We used a simplifying assumptionof 30 and 7% surface area in contact with soil for reptiles and birds,respectively, to gain some insight into relative magnitude of dermalexposure for these taxa. These are likely not representative of allreptiles and birds under all scenarios. As an example, for arborealreptiles 30% would be an overestimate, while for many snakes 30%is an underestimate.

Duration of time spent in contact with soil and the duration oftime exposed to contaminated soils are qualitatively different. Ourestimate of exposure incorporated the duration of exposure tocontaminated soils only. Contaminants are frequently heteroge-neously distributed and differential use of habitats can lead todifferent exposure durations. For example, if a lizard forages ona contaminated substrate but takes refuge in some uncontaminated

substrate, then exposure would only occur during foraging. Weassumed a 12 hour exposure duration as a way to account fordifferential habitat use in a 24 hour day. This is an importantparameter and warrants further research as exposure duration islikely to vary with species, time of day, and season. Importantly,specific behavioral activities were not included in the modelalthough it is reasonable to speculate that behavior will dramati-cally impact exposure; this again represents an understudiedelement of contaminant exposure.

Also, bioavailability is soil- and contaminant-specific as organiccontent in the soil can vary substantially among sites. For examplesoil bioavailability can vary from 0.1 to 18% for a variety of chemicals(Ehlers and Luthy, 2003). Finally, uptake across the reptileepidermis is a source of uncertainty as lipid content of theepidermis varies within and among reptile species and can havea significant impact on the uptake of both lipophilic and/orhydrophilic chemicals. Despite the assumptions and uncertainties,the results of the modeling exercise suggest that dermal exposuremay be important for reptiles and that additional research on themechanisms and magnitude of different exposure routes in reptilesis needed.

In addition to our model results, empirical evidence suggeststhat dermal exposure can be important. Although data are few,several studies have shown that dermal exposure to birds can besignificant and can perhaps exceed dietary exposures underparticular conditions (Driver et al., 1991; Henderson et al., 1994;Vyas et al., 2007). Risk models that predicted bird mortality fromfield pesticide sprays were significantly improved by inclusion ofa variable representing relative oral to dermal toxicity, which theauthor stressed should be included in future risk assessments(Mineau, 2002). Reports that quantitatively evaluated or comparedroutes of exposure for reptiles are rare. One comparison of oral anddermal toxicity of multiple chemicals to the brown tree snake(Boiga irregularis) found that both routes could cause significanttoxicity, with the oral route causing toxicity at lower concentrations(Brooks et al., 1998a). Dermal exposure caused mortality to browntree snakes exposed to 3 insecticide foggers, but no mortality wasseen in snakes exposed via inhalation (Brooks et al., 1998b). Talent(2005) exposed lizards to a pyrethrins formulation (300 mg/Lpyrethrins) including a synergist (piperonylbutoxide, 3000 mg/L)by dipping the body of the lizard in the stock solution for 2 secondsand then maintaining the lizards at multiple temperatures for 48 h.Following exposure to the formulation, mortality occurred in alltreatment groups (Talent, 2005). An enclosure study involving 2species of lizards found significant mortality from dermal exposureto deltamethrin. Dermal exposure was both directly applied to thebody, and “indirectly” acquired from contaminated soil. Fieldsurveys conducted over the 5 month period reported significantreductions in abundance (>50% reduction) for both species oflizards (Alexander et al., 2002). These results contrast with those ofLambert (1994) who reported no significant effect of deltamethrinspraying on lizard species in the field, but did not performa controlled toxicity test of deltamethrin on the lizards.

It is often considered that reptiles have a tough skin that isrelatively impermeable to water. However, skin permeability (i.e.water loss) varies among reptilian species (Gans et al., 1968), andthose with greater lipid content will have increased absorption oflipid soluble compounds, while those with lower lipid content mayhave increased absorption of polar, water-soluble compounds(Palmer, 2000). In addition, some lizards may exhibit skin perme-ability plasticity depending upon the environment to which theyare acclimated (Kattan and Lillywhite, 1989). However, Agugliaroand Reinert (2005) reported that skin permeability to water oftimber rattlesnakes from hardwood deciduous forests and lowlandcoastal wetlands did not significantly differ. Interestingly, Agugliaro

and Reinert (2005) found that skin permeability to water differedbetween juveniles and adults of the timber rattlesnake (Crota-lushorridus). Certainly, more research is required to determine therelative absorption capabilities within and among reptilian species.

5. Conclusions

We found that birds are not universally more sensitive thanreptiles to contaminants, with the exception of cholinesteraseinhibitors and sodium fluoroacetate (1080). Reptiles were moresensitive than birds to 5 out of 15 chemicals for which reliable,comparable, data could be found. However, the reptile toxicitydataset is severely lacking and as a decade ago, more toxicity dataon reptiles is needed. The use of the Western fence lizard asa toxicity testing model may prove useful for generating toxicitydata but research on other lizard species as well as turtles andsnakes would be valuable.

Avian receptors are generally expected to receive greater dietaryexposure than reptiles because endothermy is energeticallydemanding; our dietary exposure models corroborated this.However, if reptiles are more sensitive than birds for a givenchemical, the use of avian receptors as surrogates for reptiles maynot be protective. Moreover, dermal exposure could potentially bean important pathway for reptiles as indicated by the limitedempirical data (Talent, 2005; Brooks et al., 1998a,b) and our modelresults. Our dermal exposure models, by necessity, included manyassumptions but suggest that concerted research on exposure inreptiles is warranted and may reduce uncertainties associated withreptilian ERAs.

Acknowledgements

We would like to thank three anonymous reviewers forproviding constructive comments on previous drafts of themanuscript. SMWwas supported by a doctoral fellowship from theTexas Tech University office of the Provost.

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