Biomarkers of exposure and effect in a lacertid lizard (Podarcis bocagei Seoane) exposed to chlorpyrifos.

Abstract

In Europe, reptiles have been recently included in environmental risk-assessment processes for registration of plant-protection products. However, data on toxicity effects of most compounds are lacking. Chlorpyrifos is the most commonly used organophosphorus insecticide worldwide. In the present study, the authors exposed a lacertid lizard, Podarcis bocagei, to sublethal concentrations of chlorpyrifos. Individuals were exposed through spiked food for a period of 20 d (low dose 0.12 mg/kg/d, high dose 1.57 mg/kg/d). After exposure, various biomarkers of exposure and effect were evaluated, including the activities of glutathione S-transferase and enzymes involved in the glutathione redox cycle, glutathione concentrations, activities of esterases, liver and testes histopathologies, as well as locomotory and predatory behavior. The results indicate that sublethal, subchronic exposure to chlorpyrifos can affect P. bocagei in a dose-dependent manner. Adverse effects occurred at both the subindividual and individual levels, including inhibition of carboxylesterases and cholinesterases (ChEs), liver histopathological changes, and altered predatory behaviors. Animals exposed to chlorpyrifos took more time to capture and subdue prey items. The results suggest a link between effects at subindividual levels of organization with those observed at the whole individual level after exposure to environmentally realistic dosages of chlorpyrifos. Environ. Toxicol. Chem. 2012; 31: 2345-2353. © 2012 SETAC.

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BIOMARKERS OF EXPOSURE AND EFFECT IN A LACERTID LIZARD
(PODARCIS BOCAGEI SEOANE) EXPOSED TO CHLORPYRIFOS
MARIA JOSE´AMARAL,*yz JUAN C. SANCHEZ-HERNANDEZ,§ RITA C. BICHO,y MIGUEL A. CARRETERO,z RICARDO VALENTE,y
AUGUSTO M.R. FAUSTINO,k AMADEU M.V.M. SOARES,y and REINIER M. MANNy#
yCESAM and Department of Biology, University of Aveiro, Aveiro, Portugal
zCIBIO—Research Center in Biodiversity and Genetic Resources, University of Porto, Vaira ˜o, Portugal
§Laboratory of Ecotoxicology, Faculty of Environmental Sciences, University of Castilla-La Mancha, Toledo, Spain
kDepartment of Pathology and Molecular Immunology, ICBAS, University of Porto, Porto, Portugal
#Centre for Environmental Sustainability, University of Technology, Sydney, Australia
(Submitted 6 December 2011; Returned for Revision 24 January 2012; Accepted 13 June 2012)
Abstract—In Europe, reptiles have been recently included in environmental risk-assessment processes for registration of plant-
protection products. However, data on toxicity effects of most compounds are lacking. Chlorpyrifos is the most commonly used
organophosphorus insecticide worldwide. In the present study, the authors exposed a lacertid lizard, Podarcis bocagei, to sublethal
concentrations of chlorpyrifos. Individuals were exposed through spiked food for a period of 20 d (low dose 0.12mg/kg/d, high
dose 1.57mg/kg/d). After exposure, various biomarkers of exposure and effect were evaluated, including the activities of glutathione
S-transferase and enzymes involved in the glutathione redox cycle, glutathione concentrations, activities of esterases, liver and testes
histopathologies, as well as locomotory and predatory behavior. The results indicate that sublethal, subchronic exposure to chlorpyrifos
can affect P. bocagei in a dose-dependent manner. Adverse effects occurred at both the subindividual and individual levels, including
inhibition of carboxylesterases and cholinesterases (ChEs), liver histopathological changes, and altered predatory behaviors. Animals
exposed to chlorpyrifos took more time to capture and subdue prey items. The results suggest a link between effects at subindividual
levels of organization with those observed at the whole individual level after exposure to environmentally realistic dosages of
chlorpyrifos. Environ. Toxicol. Chem. 2012;31:2345–2353. # 2012 SETAC
Keywords—OrganophosphateFeeding studyReptileEcological risk assessmentTerrestrial ecotoxicology
INTRODUCTION
Globally,
pyridyl-phosphorothioate) is one of the most commonly used
organophosphorus (OP) insecticides in agricultural areas.
Chlorpyrifos tops the list of the 10 most used insecticides in
the European Union, with an annual consumption of 1,226 tons
of active substance [1]. As a member of the OP family,
chlorpyrifos inhibits the hydrolytic activity of cholinesterases
(ChEs) [2], and exposure to this pesticide is commonly eval-
uated by determination of inhibition of acetylcholinesterase
(AChE; Enzyme Commission [EC] 3.1.1.7) in nervous tissue.
Nevertheless,adverseeffectsofchlorpyrifosarenotexclusively
limited to neurotoxicity; and it has been shown to induce
oxidative stress [3] and genotoxicity [4] and to affect repro-
ductive parameters [5] and locomotor performance [6] in differ-
ent vertebrate groups. The effects of chlorpyrifos have been
extensively studied in different nontarget organisms in the
aquatic environment [7,8] and in laboratory bioassays using
rats as an animal model as well as in mammalian organs and
tissues [9]. However, the detrimental effects from chlorpyrifos
exposure in terrestrial ecosystems are less understood.
Reptiles, like other terrestrial vertebrates, respond to OP
pesticides through a demonstrable inhibition of both brain and
plasma ChE activities (S.R. Schmidt, 2003, Master’s thesis,
Texas Tech University, Lubbock, TX, USA). Several studies
chlorpyrifos(O,O-diethyl-O-3,5,6-trichloro-2-
have examined the toxic effects of OP insecticides in reptiles,
measuring ChE activities [10,11] as well as other physiological
and behavioral parameters [12,13]. Reptiles have seldom been
considered in risk-assessment processes of pesticides, and
toxicitydata fromsurrogate species suchas birdsare commonly
used to represent adverse effects in reptiles [14]. These extrap-
olations,whilenotalwaysvalid,assumethatbirdsareequallyor
more sensitive to contaminants and that contaminant exposure
is higher in birds [14]. Under the new European Commission
Regulation 1107/2009 concerning the placement of plant-
protection products on the European Union market, reptiles
would beconsideredasnontargetorganismstobe includedinthe
risk-assessment framework for plant-protection product author-
ization. However, their inclusion depends on the availability of
exposure and toxicity data, which are currently lacking.
Themodeofactionofenvironmentalcontaminantsisafocus
of predictive ecotoxicology [15]. Conceptual frameworks that
try to link effects at the level of molecular or biochemical
systems with adverse effects observed at the whole individual
level are increasingly common in predictive risk-assessment
processes[16,17].Anunderstandingofthemechanismsoftoxic
action in nontarget organisms is helpful to establish a relation-
ship between the interaction of the pesticide at the target site
and the subsequent adverse effects at the whole individual
level. Available data from the scientific literature have enabled
this mechanistic approach with regard to several classes of
environmental contaminants, such as narcotic chemicals, estro-
gen receptor agonists (e.g., nonylphenol), aryl hydrocarbon
receptor agonists (e.g., 2,3,7,8-tetrachlorodibenzo-p-dioxin,
planar polychlorinated biphenyls), and vitellogenesis disrupters
Environmental Toxicology and Chemistry, Vol. 31, No. 10, pp. 2345–2353, 2012
# 2012 SETAC
Printed in the USA
DOI: 10.1002/etc.1955
* To whom correspondence may be addressed
(mjamaral@ua.pt).
Published online 23 July 2012 in Wiley Online Library
(wileyonlinelibrary.com).
2345

Page 2

(e.g., 17b-trenbolone, prochloraz) [16]. Anticholinesterase
pesticides are applicable candidates for this approach because
ofthelarge volumeofinformationontheirmechanismofaction
as well as on their potential adverse effects on organisms.
The aim of the present study was to examine the sublethal
effects of chlorpyrifos at multiple levels of biological organ-
ization in the lizard Podarcis bocagei (Seoane, 1885). The
attempt was to establish a direct link between biochemical
processes of chlorpyrifos toxicity and detrimental effects at
the individual level using a conceptual approach named
‘‘adverse outcome pathways’’ [16,17]. In Mediterranean envi-
ronments, reptiles constitute a large component of agricultural
ecosystems [18]. Moreover, lacertid lizards have been recently
proposed as potential model species for reptile ecotoxicology
in Europe [19]. Podarcis spp., the common wall lizards, are
ubiquitous through most of central and southern Europe and
have been used in a variety of studies to assess exposure to
different contaminants such as metals [20], pesticides [21], and
polyaromatic hydrocarbons [22].
MATERIALS AND METHODS
Animal maintenance
Adult male P. bocagei were collected in the coastal dune
system of Mindelo, Vila do Conde (northwestern coast of
Portugal), in April 2010. Animals were housed individually
in glass terrariums (40?20?25cm) in a climate-controlled
room (22?18C). Terrariums contained a terra-cotta vase
(diameter, 16cm) that provided a refuge and a basking location
and a 25-W incandescent lamp that created a thermal gradient
(25–358C, 8h/d). Lighting was provided by natural sunlight,
fluorescent lighting (2?40W), and a high-pressure sodium
lamp (400W, 7,000–12,000 Lx) for 12h/d. Water was provided
in shallow dishes and renewed every 2 d. Animals were
acclimatized for 10 d before the start of the experiment and
fed daily live mealworms (Tenebrio molitor).
Pesticide exposure conditions and tissue preparation
A total of 36 animals were randomly assigned to three
experimental groups: control, low, and high dose. Nominal
concentrations of the treatment solutions were 96mg/L (low
dose) and 960mg/L (high dose). These chlorpyrifos concen-
trations were based on the application rate of the product in
agricultural areas (150/200ml/h) and a previous study that
reported chlorpyrifos residue concentrations in mealworms
(1 d, 0.53mg/kg; 5 d, 5mg/kg; and 10 d, 6.78mg/kg [23]).
Solutions were prepared daily by diluting in distilled water the
commercial product Ciclone148 EC (480g/L chlorpyrifos
with xylene and other nondefined surfactants; Sapec Agro),
covering, and agitating for 30min before administering to
mealworms using microliter syringes.
Lizardswerefedevery2dwithtwolivemealworms(2–3cm
length), previously injected with 5ml of the corresponding
chlorpyrifos solution (low and high dose) or distilled water
(control group) over a period of 20 d. At the end of the exposure
period, six individuals of each treatment were weighed, anes-
thetized by cooling on ice, killed by decapitation, and dissected.
Animal handling complied with Portuguese animal ethics
guidelines as stipulated by Direcc ¸a ˜o Geral de Veterina ´ria
and Instituto da Conservac ¸a ˜o da Natureza e Biodiversidade.
Blood was collected with a pipette from the exposed trunk and
centrifuged (800rpm for 10min at 48C) to separate serum,
which was frozen in liquid nitrogen. Liver and the left testis
wereweighed.Intestine,aportionofliver(thetwolargerlobes),
brain, and right testis were frozen in liquid nitrogen and
stored at –808C. Tissue samples were later thawed on ice
and homogenized in 0.1M Tris-HCl, 0.25M sucrose, and
1mM ethylenediaminetetraacetic acid (EDTA; pH 7.6) using
a glass-Teflon Potter-Elvehjem homogenizer. Homogenates
were centrifuged (10,000g for 20min at 48C). A 100-ml aliquot
of the supernatant (postmitochondrial fraction) was mixed with
100ml of 10% trichloroacetic acid (w/v) to remove proteins
(centrifugationat10,000g,5minat48C)andtobesubsequently
used for glutathione determination. The rest of the supernatant
was immediately frozen at –808C pending biochemical deter-
minations. Total proteins were quantified by the Bradford
method using bovine serum albumin as a standard [24]. The
left testis and the smaller liver lobe were fixed in Davidson’s
solution for 24h, then washed and stored in 70% ethanol until
histopathological analysis.
The remaining six individuals of each treatment were
subjected to a locomotor performance test and a predatory
behavioral experiment, and finally released at their collection
site after a 30-d recovery period, during which they were fed
noncontaminated mealworms.
Chlorpyrifos analyses
A set of solutions prepared identically to the test solutions
was sent foranalyses of chlorpyrifosby liquid chromatography-
tandem mass spectrometry using the QuEChERS (quick,
easy, cheap, effective, rugged, and safe) method [25]. The
assayed concentrations of these solutions were below the
nominal values anticipated: 53.4?34.3mg/L (low dose) and
699.0?129.1mg/L (high dose). Because the mixtures form
emulsions, high variability was expected.
To determine the mean dose received by each lizard, res-
idues of chlorpyrifos in a subsample of mealworms injected
with Ciclone 48 EC were also analyzed by the QuEChERS
method [25]. Mealworms were weighed and homogenized
(1:10, w/v) in 0.1M sodium phosphate buffer (pH 7.4) using
a glass-Teflon Potter-Elvehjem homogenizer. Afterward, chlor-
pyrifos was extracted by mixing 1g of homogenate with 1ml of
acetonitrile, high-performance liquid chromatography (HPLC)-
grade. The mixture was energetically shaken by hand for 1min.
Next, 0.5g NaCl was added and the mixture was shaken again
for 30s. Samples were centrifuged at 3,000g for 5min at 48C,
anda10-mlaliquotofthesupernatantwasinjectedinanAgilent
1200 HPLC system, which included a manual injector (7725i
injection valve, 20-ml loop), a vacuum degasser, a quaternary
pump, and a multiple wavelength detector. The mobile phase
consisted of H2O with 0.05% acetic acid (A) and acetonitrile
with 0.05% acetic acid (B). Chlorpyrifos was separated in an
LC-8 column (0.46cm?25cm?5mm particle size) at a flow
rate of 0.5ml/min under the following solvent program: 65% B
at 0min, increase to 95% B at 7min, and keep for 3min, then
65% B at 5min. The retention time of chlorpyrifos under
these chromatographic conditions was 9.14?0.024min. The
multiple wavelength detector was set at 290nm.
Biochemical biomarkers
Antioxidant enzymes were determined on brain, intestine,
liver, and testis. Glutathione S-transferase (GST; EC 2.5.1.18)
activity was determined spectrophotometrically at 340nm
according to Habig et al. [26]. Specific activity was expressed
as milliunits per milligram of protein using a millimolar
extinction coefficient of 9.6/mM/cm. Glutathione reductase
(GR; EC 1.6.4.2) activity was measured following the method
described by Ramos-Martinez et al. [27]. The decrease in
2346
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Page 3

absorbance at 340nm due to nicotinamide adenine dinucleotide
phosphate-oxidase (NADPH; 98%, extra pure; Sigma-Aldrich)
oxidation by reduced glutathione was measured for 1min, and
a millimolar extinction coefficient of –6.22/mM/cm was used
for specific activity calculations. All kinetics were carried out
at room temperature (20–228C) with blanks (reaction mixture
freeofsample)periodicallycheckedfornonenzymaticreaction,
and enzyme activity was corrected.
Concentrations of reduced and oxidized glutathione were
fluorimetrically determined according to the method of Hissin
and Hilf [28]. Reduced glutathione was determined by incuba-
tion of deproteinized samples in the presence of 1mg/ml ortho-
phthalaldehyde (purum, >97% HPLC-grade; Sigma-Aldrich)
in Na-phosphate (0.1M)-EDTA (5mM) buffer (pH 8.0). Deter-
mination of oxidized glutathione concentrations required a
previous step with N-ethylmaleimide (ultra, >99% HPLC-
grade; Sigma-Aldrich) to prevent glutathione oxidation; 1N
NaOH was used instead of the phosphate-EDTA buffer. In both
methods, the reaction mixture was incubated for 15min at 20 to
228C, with fluorescence read at 420nm emission and 350nm
excitation. Quantification was performed using a set of external
standards of reduced (3.27–327 nmol/ml) and oxidized (1.62–
82 nmol/ml) glutathione, which were prepared in the same way
as the samples.
Lipid peroxidation was also included as a measure of
oxidative damage. The method described by Ohkawa et al.
[29] was used to estimate lipid peroxidation. We used an
external calibration curve made with 1,1,3,3-tetramethoxypro-
pane (TMP; 99% purity; Sigma-Aldrich) to express lipid per-
oxidation as nanomoles of malondialdehyde per milligram of
protein.
We measured carboxylesterase (CbE) activity in the intes-
tine, liver, serum, and testis using two substrates: a-naphthyl
acetate (a-NA; >98% purity; Sigma-Aldrich) and 4-nitro-
phenyl valerate (4-NPV; >98% purity; Sigma-Aldrich).
Hydrolysis of a-NA was performed following the method of
Gomori and Chessick [30] as adapted by Bunyan et al. [31], by
which the formation of a-naphthol occurs in a reaction medium
that contains 25mM Tris-HCl (pH 7.6), 2mM a-NA, and the
sample. The hydrolytic reaction was stoppedafter 10min by the
addition of 2.5% (w/v) sodium dodecyl sulfate and subse-
quently 0.1% Fast Red ITR in 2.5% Triton X-100. Solutions
wereallowedtostandfor30minin thedark,andtheabsorbance
of the naphthol–fast red ITR complex was read at 530nm
(e¼33.225?103M?1cm?1). Hydrolysis of 4-NPV by CbE
activity was measured according to the method of Carr and
Chambers [32]. The reaction mixture contained 1mM 4-NPV,
50mM Tris-HCl (pH 7.5), and the sample. The reaction was
stopped after 15min using a solution of 2% (w/v) sodium
dodecyl sulfate and 2% (w/v) Tris base. The liberated 4-nitro-
phenolate was read at 405nm and quantified by a calibration
curve (5–100mM). Cholinesterase activity was determined
using two substrates:acetylthiocholine iodide (AcSCh;
>99% purity; Sigma-Aldrich) and S-butyrylthiocholine iodide
(BuSCh; >99% purity; Sigma-Aldrich) as described by Ellman
et al. [33]. Specific ChE activity was expressed as milliunits per
milligram of protein using a molar extinction coefficient of
14.15?103/M/cm. When not indicated, chemicals were pur-
chased from Scharlab.
Zymographic method
Zymographs of esterases were carried out on nondenaturing
polyacrylamide gel electrophoresis (native-PAGE) using a
Bio-Rad Tetra Cell Electrophoresis Unit. Samples (10ml) were
loaded onto 4% stacking and 12.5% resolving 0.75mm poly-
acrylamide gels and electrophoresed (25mM Tris, 192mM
glycine as running buffer) at a constant voltage of 30V for
30min and then 150V. Protein bands corresponding to CbEs
were visualized by incubation of the gel with a staining solution
containing 100mM Na-phosphate buffer (pH 6.5), 0.5mg/ml
a-NA, and 0.025g of Fast Blue RR salt. The staining solution
was prepared and filtered immediately before use. Gels were
scanned and bands individualized in a Gel Doc EZ Imager
system (Bio-Rad) and Image Lab software (Version 3.0.1,
Bio-Rad).
Histopathological analyses
Fixed samples were embedded in paraffin, and 2-mm sec-
tions were cut on a rotary microtome (Leica RM 2035). Tissues
were stained with hematoxylin and eosin and examined under a
light microscope (Olympus BX51) using an attached Olympus
camera. Liver sections were also stained with Masson’s
trichrome to assess liver fibrosis, with periodic acid-Schiff
examined for lipid or sugar accumulation, and with Perls’
Prussian Blue to verify iron pigmentation. One liver section
per individual, which compromised at least 20 fields at 400?
was scanned, and the incidence of histopathological changes
was classified through a semiquantitative scoring system:
0¼normal tissue, 1¼changes in <50% of the section, and
2¼changes in >50% of the section.
In each individual testis section, spermatogenesis develop-
ment was assessed in 20 impartially chosen seminiferous
tubules. Several morphological parameters were measured
using image analysis software (ImageJ, U.S. National Institutes
of Health). Tubular diameter and, along this same axis, the
width of the Sertoli cell layer and the width of each spermato-
genesis phase (spermatogonia, spermatocytes I and II, sperma-
tozoa) were recorded.
Behavioral performance
Locomotor performance was evaluated by measuring max-
imalsprintspeedonahorizontalcorksubstrate(2?0.1msprint
track). Each lizard was fasted for 48h before the start of the
experiment, weighed, warmed to its optimal temperature of
338C (M.J. Amaral, CIBIO and University of Porto, Portugal,
unpublished data), and raced three times, with at least a 1-h rest
interval between trials. Lizards were hand-chased through the
track,andtrialswererecordedwithavideocamera(Sony/DCR-
HC46, 25 fps). Following Holem et al. [34], mean maximum
speed was calculated as the average of the fastest 0.20-m
interval in each of the three replicate sprints and maximum
speed as the fastest 0.20-m interval. Trials were repeated when
lizards did not run continuously.
Predatory behavior was assessed in fasted (96h) individuals
tostimulatepreyattack.Eachindividualwasplacedinanempty
glass terrarium (40?20?25cm) at optimal temperature
15min before the start of the experiment for acclimation
purposes. Test terrariums were visually isolated. At commence-
ment of the trial, a mealworm (1.5cm mean length) was
introduced. The behavior of each lizard was recorded with a
camera (Sony/DCR-HC46, 25 fps) placed on top of the terra-
rium. The trial was terminated if the mealworm was not
consumed within 15min. Videos were analyzed with video
frame capture software (FrameShots, EOF Productions),
recording the time each individual took to attack the mealworm
(time of latency to attack) and the time each took to subdue and
swallow the prey (manipulation time).
Chlorpyrifos effects on a lacertid lizard
Environ. Toxicol. Chem. 31, 20122347

Page 4

Statistical analyses
The results obtained for each biomarker from the treatment
groups were compared to the control using analyses of variance
(ANOVA), analyses of covariance (ANCOVA), multiple anal-
yses of covariance (MANCOVA), and/or repeated measures
analysis of variance followed by a Dunnett’s multiple compar-
ison post hoc test. Treatment was the dependent variable, and
snout–vent length and body mass were used as covariates to
check significant interactions. When the assumptions for anal-
ysis of variance were not met (i.e., homogeneity of variance,
normality), we used the nonparametric Kruskal-Wallis test to
compare data. Behavioral data were ranked prior to the anal-
yses. Differences in the prevalence and intensity of histological
changes in liver between treatments were compared with the
control using Pearson’s chi-squared test. Level of significance
was a¼0.05.
RESULTS
During the entire period of the experiment, animals were
exposed on each feeding day to 0.53?0.34 and 6.99?1.29mg
of chlorpyrifos per lizard in the low- and high-dose groups,
respectively (data not shown). This is equivalent to a mean
chlorpyrifos exposure of 0.12 (0.05–0.17) mg/kg/d in the low-
dose group and 1.57 (1.46–1.65) mg/kg/d in the high-dose
group every other day during a period of 20 d. The dosages
in mealworms were within the expected values of 2.38mg of
chlorpyrifos/gofmealworminthelow-dosegroupand23.68mg
of chlorpyrifos/g of mealworm in the high-dose group (1g of
mealworms corresponds to 10–12 individuals). No animals died
or showed any visible symptoms of cholinergic poisoning
during the exposure period.
Biochemical biomarkers
In general, exposure to chlorpyrifos did not cause any
significant variation in the activity of the antioxidant enzymes
(GSTandGR),lipidperoxidation,orglutathioneconcentrations
(reduced glutathione [GSH] and oxidized glutathione [GSSG])
in the different tissues examined (Table 1). The results differed
on a tissue basis, and in general, there was high individual
variability. For example, GST activity was highest in liver,
followed by testes, intestine, and brain. Whereas GR activity
was more stable between treatments and had similar levels in
intestine,liver,andtestis,andwas10timeslowerinbrain.Lipid
peroxidation was higher in liver and intestine than in testis and
showed a slight increase with exposure in intestine and testis.
The only significant difference we detected in the oxidative
stress biomarkers was in GSH concentrations in brain after
exposure to the low dose of chlorpyrifos (ANOVA treatment,
F2,14¼5.2, p¼0.02). Higher concentrations of GSH also
occurred in intestine, and there was a slight increase in GSH
content in testes with chlorpyrifos exposure; but in all cases
there was high individual variability, and GSSG levels
remained similar between treatments. In both brain and intes-
tine, the concentrations of GSH and GSSG were similar and the
ratio of GSH to GSSG was close to one. Total glutathione
content was higher in liver when compared with other tissues,
whereas the ratio of GSH to GSSG was higher in testes.
Activity of CbE was recorded in all tissues and decreased in
a dose-dependent manner with exposure to chlorpyrifos (Figs. 1
and 2a). Esterase activity was tissue (intestine, liver, serum,
testis) and substrate (a-NA and 4-NPV) specific. Statistically
significant differences in enzyme activities were detected in all
tissues except testis. Significant differences in CbE activity
using a-NA (Fig. 1a) were detected in the two treatment groups
whenintestine(ANOVAtreatment,F2,14¼22.9,p<0.001)and
liver (ANOVA treatment, F2,14¼45.7, p<0.001) were used as
esterase sources. In serum (Fig. 2a), CbE activity was inhibited
in lizards exposed to the higher dose (Kruskal-Wallis treatment,
H2,17¼11.1, p¼0.004). With the a-NA substrate, CbE activ-
ities were inhibited by 56/85%, 50/81%, and 47/51% (low dose/
high dose)in intestine, liver,and testis,respectively. In serum,a
slight induction was found in the low-dose lizards, 112%,
whereas an 86% inhibition was found in the high-dose group.
The highest activity using a-NA was found in serum, followed
by intestine. Similarly, CbE activity with the 4-NPV substrate
(Figs. 1b and 2a) was also significantly inhibited in both treat-
ment groups in liver (ANOVA treatment, F2,14¼12.6,
p<0.001). In intestine and serum, inhibition was significant
depressed only in the high-dose group (intestine ANOVA
treatment, F2,14¼9.1, p¼0.003; serum Kruskal-Wallis treat-
ment, H2,17¼11.1, p¼0.004). Levels of CbE inhibition with
4-NPV substrate were 55/87%, 57/80%, and 34/54% in intes-
tine, liver, and testis, respectively. The highest CbE activities
with 4-NPV were found in serum, followed by liver. In general,
CbE activities were twice as high with the 4-NPV substrate.
Table 1. Oxidative stress parameters in Podarcis bocagei after a 20-d exposure to chlorpyrifosa
Biomarker
n
GST
(mU/mg protein)
GR
(mU/mg protein)
GSH
(ng/mg protein)
GSSG
(ng/mg protein)
tGS
(ng/mg protein)GSH/GSSG
LPO
(ng MDA/mg protein)
Brain Control
Low dose
High dose
Control
Low dose
High dose
Control
Low dose
High dose
Control
Low dose
High dose
6
5
6
6
6
5
6
5
6
6
6
6
31.5?3.2
78.9?86.8
44.7?14.5
374.2?137.0
268.4?130.8
358.0?271.2
1,428.0?290.7
1,748.7?553.3
1,277.0?396.7
607.4?352.9
567.2?119.0
546.6?137.6
5.0?0.8
5.4?1.5
4.84?2.4
77.8?18.1
78.2?35.5
75.2?16.7
63.4?12.2
83.3?49.0
61.6?7.2
49.0?30.5
53.5?34.1
53.5?29.4
17.3?3.5b
22.0?12.7b
18.3?4.8
18.8?6.6
23.4?15.1
12.0?5.1
81.7?14.6
61.3?40.2
64.8?27.6
41.9?22.8
58.8?35.8
66.6?34.5
23.0?13.3
29.9?28.0
15.0?12.8
21.5?7.9
23.2?5.3
18.8?4.3
39.9?23.5
39.8?14.0
26.8?5.3
10.9?9.0
8.8?4.8
10.3?4.7
40.2?16.1
51.9?39.1
33.2?15.6
40.2?10.1
46.7?16.4
30.8?7.8
121.6?28.8
101.1?35.5
91.6?25.2
57.8?21.3
67.7?39.5
76.9?34.7
1.1?0.8
1.2?1.3
1.9?1.4
1.0?0.5
1.0?0.6
0.6?0.3
2.8?1.9
1.8?1.4
2.6?1.2
5.4?3.2
6.7?4.2
7.3?4.6
Intestine 4,511.3?2,669.3
6,132.8?3,122.1
7,475.8?2,298.3
9,457.7?2,633.2
12,481.5?6,905.0
8,809?1,821.8
8,668.8?2,008.4
12,237.5?8,899.4
12,812.1?3,439.4
Liver
Testis
aControl (0mg/kg/d body wt), low dose (0.12mg/kg/d body wt), and high dose (1.57mg/kg/d body wt). All parameters are presented as means?standard
deviations.
bDatathataresignificantlydifferentfromcontrolsfollowingDunnett’sposthocresultsafteranalysisofvariance(ANOVA)orposthocpairedcomparisonsafter
Kruskal-Wallis, p<0.05.
GST¼glutathione S-transferase; GR¼glutathione reductase; GSH¼reduced glutathione; GSSG¼oxidized glutathione; tGS¼total glutathione; LPO¼lipid
peroxidation; MDA¼malondialdehyde.
2348
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Page 5

We distinguished multiple CbE isoforms by native PAGE,
which was tissue-specific (Fig. 1c). Liver had the higher
abundance of CbE isoforms (nine protein bands). Furthermore,
the relative mobility of most of the stained bands was similar
among tissues, which suggests the occurrence of the same
proteins displaying CbE activity in the selected tissues. The
staining intensity of CbE bands decreased in lizards exposed to
chlorpyrifos irrespective of the tissue.
The use of two selective substrates allowed us to distinguish
different ChEs in the analyzed tissues. Marked variations were
noted on a tissue basis, and ChE activity was significantly
inhibited in the high-dose group in all tissues (Figs. 2b
and 3). In brain, ChE activity was slightly induced in the
low-dose group, 123% of control activity, and inhibited by
70% in the high-dose group (ANOVA treatment, F2,14¼16.2,
p<0.001). Intestinal ChE activity was induced by 149% and
significantlyinhibitedby82%inthelow-andhigh-dosegroups,
respectively (ANOVA treatment, F2,13¼48.2, p<0.001).
Similarly, serum ChE activity toward AcSCh was induced
by 112% and significantly inhibited by 82% in the low- and
high-dose groups (ANOVA treatment, F2,14¼26.6, p<0.001),
respectively. However, hydrolysis of the substrate BuSCh by
serum ChE activity was significantly inhibited in both treat-
ments, by 57 and 96%, respectively (ANOVA treatment,
F2,13¼52.7, p<0.001).
Histopathological analyses
Livers from control individuals were characterized by
normal tissue with some signs of congestion (three out of six
Fig. 2. Esterase activity in serum in Podarcis bocagei exposed to sublethal
concentrations of chlorpyrifos for a period of 20 d. Control (Ctr) (0mg/kg/d
body wt), low dose (0.12mg/kg/d body wt), and high dose (1.57mg/kg/d
body wt). Esterase activity was measured using the substrates a-naphthyl
acetate (a-NA), 4-nitrophenyl acetate (4-NPV), acetylthiocholine iodide
(AcSCh),andS-butyrylthiocholineiodide(BuSCh).n¼6,exceptinlowdose
(n¼5). Asterisks indicate statistically significant results of Dunnett’s post
hoctestsafteranalysisofvariance(ANOVA)orposthocpairedcomparisons
afterKruskal-Wallis:?p<0.05,??p<0.01,???p<0.001.Dataareexpressed
as mean?standard error.
Fig. 3. Cholinesterase activity in brain and intestine in Podarcis bocagei
exposed to sublethal concentrations of chlorpyrifos for a period of 20 d.
Control (Ctr) (0mg/kg/d body wt), low dose (0.12mg/kg/d body wt), and
high dose (1.57mg/kg/d body wt). Cholinesterase activity was measured
in brain using the substrate acetylthiocholine iodide and in intestine using
S-butyrylthiocholine iodide. n¼6, except in intestine-high dose (n¼5).
AsterisksrepresentsignificantdifferencesfollowingDunnett’sposthoctests:
???p<0.001. Data are expressed as mean?standard error.
Fig. 1. Carboxylesterase(CbE)activityinintestine,liver,andtestisinPodarcisbocageiexposedtosublethalconcentrationsofchlorpyrifosforaperiodof20d.
Control (Ctr) (0mg/kg/d body wt), low dose (0.12mg/kg/dbody wt), and high dose (1.57mg/kg/dbody wt). Esterase activity was measured using the substrates
(a)a-naphthylacetate(a-NA)and(b)4-nitrophenylacetate(4-NPV).n¼6,exceptinlowdose(liverandserum)andhighdose(intestine),wheren¼5.Asterisks
areDunnett’sposthocresultsafteranalysisofvariance(ANOVA)orposthocpairedcomparisonsafterKruskal-Wallis:??p<0.01,???p<0.001.(c)Zymographof
CbEactivityinintestine,testis,intestinewhere.1,4,and7arecontrol;2,5,and8arehigh-dose;and3,6,and9arelow-dose.Dataareexpressedasmean?standard
error. [Color figure can be seen in the online version of this article, available at wileyonlinelibrary.com]
Chlorpyrifos effects on a lacertid lizard
Environ. Toxicol. Chem. 31, 2012 2349

Page 6

individuals, Fig. 4a). One of these individuals presented hep-
atocyte degeneration, vacuolation, and fibrosis. Livers of indi-
viduals from both treatment groups consistently presented
congestion and hepatocyte vacuolation. In general, lizards in
the high-dose group presented a higher prevalence of histo-
logical changes, including severe congestion, hepatocyte
vacuolation, and light fibrosis (Fig. 4b and c). In the high-dose
treatment, the incidence of fibrosis, hepatocyte degeneration,
and vacuolation approached significance when compared with
the control (Pearson chi-squared tests, df¼2: fibrosis x2¼5.1,
p¼0.08; congestion x2¼9.1, p¼0.01; hepatocyte degenera-
tion x2¼4.4, p¼0.11; Pearson chi-squared tests, df¼4:
hepatocyte vacuolization x2¼6.6, p¼0.15).
Testes histology presented no alterations in morphology,
irrespective of exposure dose (data not shown). In both control
and treated groups, normal testes with full spermatogenic
activity were observed. Seminiferous tubule diameter was
not different in any treatment group when compared to control
animals (ANCOVA treatment, F2,14¼0.04, p¼0.96). Adition-
ally, no size differences were detected for the Sertoli cell layer
(Kruskal-Wallis treatment, H2,18¼1.7, p¼0.42) or for the
layers of the different phases of spermatogenesis: spermatogo-
nias (ANCOVA treatment, F2,14¼0.4, p¼0.70), spermatocyte I
(ANCOVA treatment, F2,14¼0.05, p¼1.0), spermatocyte II
(ANCOVA treatment, F2,14¼0.2, p¼0.84), and spermatozoa
(ANCOVA treatment, F2,14¼0.1, p¼0.90).
Behavioral performance
Mean maximum speed and maximum speed were not
affected by exposure to chlorpyrifos (ANCOVA treatment
mean maximum speed, F2,14¼1.7, p¼0.22; ANCOVA treat-
ment maximum speed, F2,14¼0.5, p¼0.64). In the predatory
behaviorexperiment,onlyoneofthe18lizardsdidnotconsume
the prey item in the 15-min period. Considering time of
latency to attack, an outlier was removed from the data
(value>mean?2 SD), and no difference between exposures
was found (ANCOVA treatment time of latency to attack,
F2,12¼1.5, p¼0.27). For manipulation time our results,
although not statistically significant, suggest that animals
exposed to chlorpyrifos take more time to manipulate the
prey (ANCOVA treatment manipulation time, F2,13¼3.1,
p¼0.08). Considering both variables together, lizards in the
high-dose treatment took significantly more time to subdue and
swallow the prey than lizards from the control group (Fig. 5)
(MANCOVA treatment, Wilks F4,22¼2.8, p¼0.048).
DISCUSSION
Reptiles inhabiting agroecosystems can be exposed to pes-
ticides through the ingestion of contaminated food. However,
detection of pesticide exposure in the field is usually complex
becausetherearealwaysmultipleandfluctuatingbiologicaland
environmental variables that impact pesticide uptake and sub-
lethal effects. This limitation of using field data is compounded
when the understanding of the biochemistry and physiology of
theorganismofinterest(e.g.,lizards)islimited.Whenexposure
results in toxic effects, they are usually sublethal in nature; and
thus, their occurrence under natural conditions is difficult to
resolve because they can be masked by other nontoxicological
variables. Laboratory studies sacrifice ecological realism but
allow a higher degree of control over exposure parameters,
allowing detection of those sublethal effects. Our results sug-
gest that sublethal, subchronic exposure to chlorpyrifos can
affect P. bocagei in a dose-dependent manner. Adverse effects
of ecologically realistic chlorpyrifos doses occurred at both the
subindividual and individual levels, including differences in
esterase activities, liver histopathological changes, and altered
predatory behaviors.
Glutathione-dependent enzymes, which are involved in
cellular antioxidant defense, have been studied in reptiles,
mainly to assess the response to overwintering and hypoxic
conditions (reviewed in Mitchelmore et al. [35]). Induction of
glutathione peroxidase after exposure to cadmium was reported
for adult male Uromastyx aegyptius lizards [36]. Nevertheless,
glutathione-dependent enzymes have the potential to limit the
damage caused by reactive oxygen species generated during
Fig. 4. Representative liver sections of Podarcis bocagei exposed to sublethal concentrations of chlorpyrifos for a period of 20 d. (a) Liver section of a control
animalrepresentinganormalliver(hematoxylinandeosin400?).(bandc)Examplesofliversectionsofindividualsexposedtothehighdose(699.0?129.1mg/
L).(b)Liversectionwithhepatocytevacuolation(arrows,hematoxylinandeosin400?).(c)Liversectionwithcongestion(arrow,hematoxylinandeosin400?).
[Color figure can be seen in the online version of this article, available at wileyonlinelibrary.com]
Fig. 5. Predatory behavior, assessed by time of latency to attack (TLA) and
manipulation time (Man), in Podarcis bocagei exposed to sublethal
concentrations of chlorpyrifos for a period of 20 d. Ctr (0mg/kg/d body
wt),lowdose(0.12mg/kg/dbodywt),andhighdose(1.57mg/kg/dbodywt).
Bars represent meansþSE (n¼6, except for low dose n¼5).
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Page 7

pesticide-detoxificationprocesses andhavebeenusedinseveral
studies to assess pesticide exposure and effects (reviewed for
fishinvanderOostetal.[37]).Lipidperoxidationhasbeenused
as a quantitative measure of the extent of oxidative damage
caused by contaminants. In the present study, we found no signs
of oxidative stress caused by chlorpyrifos exposure. This is
contrary to the results of a study in rats, for which a dose of
6.75mg/kg body weight for a period of 28 d resulted in a
significant increase in serum lipid peroxidation and decreased
significantly the activity of serum GST [38]. Another rat study
also reported accumulation of malondialdehyde in different
tissues and a dose-dependent decrease in antioxidant enzymes
after 1 d of exposure to 38.8mg/kg body weight [3]. In both
studies, rats were orally exposed to a higher dosage of chlor-
pyrifos than the higher one in the present study.
Chlorpyrifos exposure resulted in the inhibition of CbEs
in all tissues analyzed in a dose-dependent manner, except for
serum CbE, which was strongly depressed at the higher dose of
the OP. Inhibition of CbE activity has been used as a biomarker
of OP exposure in different species, and several authors have
suggested that CbEs can have a protective role by binding
stoichiometrically to OP pesticides, decreasing OP concentra-
tionandtoxicity[39,40].Thus,effectsofpesticideexposurecan
bealleviatedindifferenttissuesbythelevelsofCbEactivity,its
sensitivity to chlorpyrifos inhibition, and the presence of multi-
ple isozymes [39]. Different studies have also suggested that
CbE activity could be a more sensitive biomarker than brain
cholinesterase activity, the most commonly used biomarker of
OP exposure [41]. It is widely accepted that CbE activity
provides a protective effect against OP exposure because of
its higher sensitivity to inhibition by these agrochemicals
compared to ChE activity [41]. Results in the present study
support the aforementioned effect. Animals exposed to the
lower dose of chlorpyrifos showed significant inhibition of
CbE activity in the liver and intestine, whereas this depression
was not found in the brain and intestine ChE activities (Figs. 1
and 3). Sanchez-Hernandez and Wheelock [40] suggested that
intestinal CbEs would have an important role in OP detoxifi-
cation as one of the first surfaces of contact for feeding
exposure. These results reinforce the need to study different
tissues and substrates as different isozymes seem to be involved
in the hydrolysis of the different substrates in multiple tissues.
The only tissue where no significant differences were found for
CbE activity was the testis, although a general trend of activity
depression with increased dose was observed. Several studies
with mammals and other organisms have demonstrated that
CbEs have an important role in reproduction by altering the
metabolic routes of testosterone production [5]. In the present
study,enzymaticactivitywasdependentontissueandsubstrate.
TheprotectiveroleofCbEswasnotclearinserumCbEactivity,
but it was apparent in both liver and intestine CbE activities.
Liver CbEs were strongly inhibited after exposure to both doses
of chlorpyrifos. Similarly, intestinal CbE activity was also
depressed. However, we detected higher intestinal CbE activity
in control than in liver with the a-Na substrate.
Cholinesterases, in particular AChE, have been intensively
studied as biomarkers of pesticide exposure and effect, in
particular for OP and carbamate pesticides. In reptile brains,
AChE is expected to represent >90% of total ChEs (S.R.
Schmidt, 2003, Master’s thesis, Texas Tech University,
Lubbock,TX,USA).TherelationshipbetweenAChEinhibition
and death has been intensively debated. Reductions of 50 to
80%inAChEactivityhavebeenrelatedtomortalityindifferent
vertebrate species (e.g., in birds [42]). In the present study,
inhibition levels of 70% of brain ChE activity recorded in the
group exposed to the high dose did not result in the death of
any individual. These levels of inhibition are similar to those
described for the lizard Gallotia gallotia after acute exposure to
parathion, where ChE inactivation levels of 84 and 70% were
observed without mortality [43].
Even if not lethal, chlorpyrifos can interfere with brain ChEs
and the activity of the central nervous system, impairing
essential behavioral functions, like predator–prey interactions
[44]. Cholinesterases in other tissues, such as blood, have been
suggested to have, like CbEs, a protective role over brain ChE,
detoxifying OPs before they reach the brain [41]. Different
studies in lizards have further compared the levels of ChE in
those tissues (mainly serum) and found them to be correlated
with those of brain ChE [43]. In the present study, this relation-
ship could only be established with serum ChE (assessed with
BuSCh substrate). In reptiles, 80% of the total ChE activity in
serum is attributable to BChE activity [11,13]. Serum ChE
(BuSCh substrate) was inhibited at the low dose before any
effect was observed in brain ChE. Cholinesterase in intestine
and serum (AcSCh substrate) was not inhibited at the low dose.
At the high dose, ChEs were strongly inhibited in all tissues, so
anyprotectiverolewouldbelost.Sanchez-Hernandez etal.[43]
obtained similar results to the findings presented here, which
show an exponential decay ofAChE activity when serum BChE
reached 90% of inhibition.
We detected a higher prevalence of liver histopathological
changes in animals exposed to chlorpyrifos. These changes
occurred mainly in the high-dose group and included fibrosis,
hepatocyte degeneration, and vacuolation. Despite being stat-
istically nonsignificant, these alterations might be indicative
of metabolic stress. Studies in rats have demonstrated a link
between exposure to chlorpyrifos and the appearance of similar
lesions [38,45]. Toxic effects of chlorpyrifos on testicular
function have also been noted in rats [5,9]. Nevertheless, in
the present study, no effects were noted. Normal testicular
histology with all the successive stages of spermatogenesis
was observed in the three groups.
If exposure to pesticides results in behavioral alterations,
growth, reproductive success, and survival might be affected.
Locomotor performance is probably the individual parameter
that has been employed most in studies with reptiles exposed to
insecticides. Nevertheless, these studies seem to indicate that
this flight response is not a parameter sufficiently sensitive to
address pesticide exposure [46,47]. In the present study, the
results similarly show this test to be an insensitive biomarker of
chlorpyrifos exposure, with no difference between control and
treated animals. In contrast,predatory behaviors provide a more
sensitive indicator of locomotor impairment because they
involve multiple neuromuscular systems, rather than simply
a sympathetically controlled, reflexive flight response. Bain
et al. [13] observed that animals exposed to a high dose of
fenitrothion tended to make more attempts to catch prey.
The authors speculated that this type of response could be
related to decreased visual acuity, changes in muscle activity,
or decreased muscle coordination [13]. We found that after
exposure to chlorpyrifos, animals took more time to capture and
swallow a prey item, despite prior habituation to the prey.
Inhibition of ChEs, which was demonstrated in the same
animals, provides a mechanistic explanation. Inhibition of AChE
causes accumulation of acetylcholine in synaptic junctions. The
accumulation of acetylcholine can alter neuromuscular abilities
and, thus, decrease the muscle coordination essential for the
manipulation of prey.
Chlorpyrifos effects on a lacertid lizard
Environ. Toxicol. Chem. 31, 20122351

Page 8

Wecanpresumethatanimalswillrecoverwhentheexposure
terminates. However, if we take into consideration the slow rate
of esterase recovery that has been reported for reptiles [13,43],
these results can represent a high exposure risk. Animals would
be impaired not only during the exposure period but also for a
significant period afterward. Insecticide application usually
occurs during spring and autumn, the period when animals
are acquiring energy for production and reproduction and to
survive the winter. To fulfill energetic requirements, compro-
mised individuals would have to spend more time foraging to
catch the same amount of prey. During foraging, lizards are
potentially at a higher risk of predation. Thus, we can speculate
that if an animal compensates for deficiencies in predatory
ability by foraging for longer periods, it is increasing its risk of
predation. On the other hand, if an animal does not compensate,
then reduced feeding rates will have implications for its
energetic balance.
Our exposure rates were under the median lethal concen-
tration (LC50) values reported for several vertebrate species
including birds (5–157mg/kg body wt) and mammals (151–
1,000mg/kg body wt) [48]. For amphibians, the reported doses
are given as concentration in the aquatic medium and not easily
comparable, ranging from 2.4 to 66.2mg/L, depending on the
species [48]. Given the lack of data, it is difficult to compare the
toxicity of chlorpyrifos to reptiles and other terrestrial verte-
brates and, thus, evaluate if birds can be used in risk-assessment
processes as surrogate species.
CONCLUSIONS
In the present study, we demonstrated that a subchronic
exposure to environmentally relevant doses of chlorpyrifos can
inhibit esterases and impair predatory behaviors. It remains to
beseenifthebiochemicalandbehavioraleffectsobservedinthe
present study are likely to translate into any short- or long-term
implications for populations. Further work in the form of long-
term mesocosm experiments will be needed to examine these
issues.
Acknowledgement—WeappreciatetheassistanceofA.Kaliuntzopoulou,N.
Sillero,andS.Freitasinanimalcollection.Allanimalswerecollectedundera
permitissuedbythe InstitutodaConservac ¸a ˜odaNaturezae Biodiversidade.
This research was supported by FEDER through COMPETE-Programa
Operacional Factores de Competitividade and by national funding through
Fundac ¸a ˜o para a Cie ˆncia e Tecnologia (FCT), within the research project
LacertidLizardsasBioindicatorsofPesticideExposureandToxicity(LAB-
PET)inintensivemarketgardenagriculture(FCTPTDC/AMB/64497/2006)
and through an FCT PhD grant to M. J. Amaral (SFRH/BD/31470/2006).
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