Chemosphere 87 (2012) 757–764
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The use of a lacertid lizard as a model for reptile ecotoxicology studies - Part 1
Field demographics and morphology
Maria José Amaral a,b,⇑, Miguel A. Carretero b, Rita C. Bicho a, Amadeu M.V.M. Soares a, Reinier M. Mann a,c,d
a
CESAM & Departmento de Biologia, Universidade de Aveiro, 3810-193 Aveiro, Portugal
CIBIO – Centro de Investigação em Biodiversidade e Recursos Genéticos, Universidade do Porto, Campus Agrário de Vairão, 4485-661 Vairão, Portugal
c
Centre for Environmental Sustainability, University of Technology – Sydney, Broadway, NSW 2007, Australia
d
Hydrobiology, Auchenflower, Queensland 4066, Australia
b
a r t i c l e
i n f o
Article history:
Received 6 October 2011
Received in revised form 25 December 2011
Accepted 30 December 2011
Available online 8 February 2012
Keywords:
Field study
Population
Podarcis bocagei
Environmental contamination
Reptile ecotoxicology
a b s t r a c t
At the European level, lacertid lizards have been proposed as potential model species for reptile ecotoxicology. We studied demographic and morphological aspects of natural field subpopulations of Podarcis
bocagei inhabiting similar agricultural habitats which were either regularly exposed to pesticides, or
not. Parameters examined in this study included population size and density, sex ratio, adult body size,
fluctuating asymmetry in femoral pores and parasite prevalence. In general, we detected few statistically
significant differences between the exposed and reference subpopulations. Although field situations are
ecologically complex and factors other than pesticides may be acting, the absence of observable effects on
field subpopulations is probably indicative that lizards are coping or compensating for this level of
exposure.
Ó 2012 Elsevier Ltd. All rights reserved.
1. Introduction
In 2010, 28% of the reptiles evaluated by the International
Union for the Conservation of Nature (IUCN) were listed as
Critically Endangered (CR), Endangered (EN) or Vulnerable (VU)
(IUCN, 2010) and environmental pollution has been recognized
as one of the main contributing factors (Gibbons et al., 2000; Lange
et al., 2009; Todd et al., 2010). Despite consistent calls for greater
emphasis on reptile ecotoxicology research, there is still a lack of
knowledge regarding the responses of reptiles to contaminants
(Hall, 1980; Hopkins, 2000; Sparling et al., 2010). Pesticides, in
view of their biocide activity, scale and volume of application,
are the major contaminant of concern for terrestrial environments
(Sánchez-Bayo, 2011). Reptiles, as non target organisms, can be
directly exposed to pesticides through various routes, including:
(a) ingestion of contaminated food; (b) accidental or deliberate
ingestion of soil; (c) inhalation; (d) maternal transfer to eggs/
young; (e) dermal exposure; (f) absorption by eggs of contaminants from surrounding environments (Lambert, 1997; Rauschenberger et al., 2004; e.g. de Solla et al., 2006; Rich and Talent, 2009;
Weir et al., 2010) and indirectly affected as a consequence of
diminished food availability or plant cover (Sánchez-Bayo, 2011).
⇑ Corresponding author at: CESAM & Departmento de Biologia, Universidade de
Aveiro, 3810-193 Aveiro, Portugal. Tel.: +351 234370790; fax: +351 234372587.
E-mail address: mjamaral@ua.pt (M.J. Amaral).
0045-6535/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved.
doi:10.1016/j.chemosphere.2011.12.075
Modern pesticides have, over the years, become more specific
and safer for non target organisms. Registration of a pesticide for
use in agriculture usually requires the submission of data to a
regulatory agency (in the EU, the European Food Safety Authority
through the Council Directive EC 1107/2009). These agencies generally require evidence that the chemical is not overly toxic to
aquatic fauna, bees, terrestrial mammals and birds. Information
on toxicity to reptiles is only required if it is available. As a consequence, hundreds of chemicals are continuously being registered
for use in agricultural areas worldwide, and the impact of these
chemicals on reptiles remains largely unknown.
In all fields of toxicology, specific species are being adopted to
act as surrogate models for field and laboratory based research.
Given their prevalence in Europe, particularly in the Mediterranean
basin (Arnold and Ovenden, 2002; Carretero, 2004), lacertid lizards
were identified as potential model species for reptile ecotoxicology
(e.g. Mann et al., 2006; Cardone et al., 2008). Primarily, the major
advantage of these taxa is one of abundance; in suitable habitat
types, these animals occur at very high densities compared to other
vertebrate taxa, for example 10 lizards to 1 mammal (Valverde,
1967). Moreover, adult lacertids are characterized by strong site
fidelity, spending their entire lives within small home ranges and
are usually generalist predators (Galán, 1999; Carretero, 2004).
Therefore, their exposure history can be assumed with some certainty. Also, they have relatively short life spans (3–4 years). Hence,
they are expected to provide inference on short term impacts.
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M.J. Amaral et al. / Chemosphere 87 (2012) 757–764
Lacertid lizards are also relatively easy to maintain in captivity.
Most importantly, like many small opportunistic lizard species
around the world, they have taken advantage of man-made habitats
(Graziani et al., 2006) that often allow many species to persist and
even thrive in otherwise harsh environments. The existence of
comparable habitats that occur over wide geographic regions, such
as stone walls in Mediterranean agriculture zones, allows for a comparison between areas where pesticides are applied on a regular
basis and areas where no pesticides are in use.
At the population level, the effects of exposure to contaminants
have rarely been confirmed, although it is expected that they will
influence subpopulations in subtle ways in conjunction with other
biotic and abiotic factors like disease, predation, competition, food
availability, climatic conditions and habitat quality (Banks and
Stark, 1998; Meyers-Schone, 2000). At least two studies have related declines in reptile subpopulations with sublethal contaminant exposures through modifications on population dynamics
(Guillette et al., 1994; Shelby and Mendonca, 2001). Moreover,
numerous parameters of exposure and effect in subpopulations
and individuals have been identified as potentially useful, although
very few of these have been applied reliably to reptile subpopulations to document population declines.
The primary hypothesis for this study is that the physiological
effects of exposure to a cocktail of pesticides will be manifested
within the exposed populations of Podarcis bocagei, either as
changes in population structure, or as changes in individual biomarkers of exposure and effect. Responses at higher levels of organization are usually hard to relate with contaminant exposure.
Although metrics of exposure at the individual level have been
used to predict effects, they have seldom been associated with effects at community or population levels (Rattner et al., 2000). We
adopted a tiered approach that examined subpopulations of the
lacertid lizard P. bocagei living in similar habitats (agricultural
stone walls), that were either regularly exposed to pesticides, or
not. The study has been separated into two subsets of data: 1.
demographics and morphological aspects of natural field subpopulations, the details of which are described here; and 2. biomarkers
of exposure and toxicity at physiological, biochemical and cellular
levels, which are presented in a companion paper (Amaral et al.,
2012).
2. Methods
2.1. Study area and species
Fieldwork was carried out in six localities in the north coastal
region of Portugal (Fig. 1a). These sites were all agricultural fields
with adjoining non cropped field margins and stone walls that provided habitat and shelter for reptile species (e.g. Fig. 1b). Four sites
(Exp 1, Exp 2, Exp 3, Exp 4) were under intensive agriculture (exposed) in Vila do Conde (41°190 N, 8°400 W) and two sites (Ref 1,
Ref 2) were under organic agriculture (reference) in Vila Nova de
Famalicão (41°260 N, 8°300 W), approximately 20 km further east.
The exposed sites were corn fields (Zea mays L.) with a winter rotation of ryegrass (Lolium multiflorum Lam.), where fertilizers, herbicides and insecticides have been applied routinely for more than
30 years. The two reference sites included a pasture and an organic
farming field (several cultures) with no history of chemical application. The organic farming fields were within 500 m of conventional sites that are routinely sprayed, therefore some exposure
may occur from drift but it was predicted to be quite low. The sites
were characterized through consultation with the Cooperativa Agrícola de Vila do Conde regarding agricultural practices in the region
and with landowners concerning current and historical land and
pesticide usage. In the exposed sites, farmers are mainly using a
Fig. 1a. Location of the sampling sites on NW Portugal. Exposed sites are shown by
a dark or gray circle and the reference areas by a white circle.
Fig. 1b. Example of a sampling site (Exp 2).
combination of herbicides: a corn pre seedling herbicide, which
is usually complemented by a second application of a post emergence herbicide. A third herbicide is used by some of the farmers
to control plants in field margins, while an insecticide application
is only carried out when necessary. The products used are usually
changed every 2–3 years to avoid weed resistance or because of EU
restrictions.
The six sites are included in the Temperate Oceanic Sub-Mediterranean climate type with an average temperature <15 °C, a precipitation >1000 mm and a high relative humidity (Rivas-Martinez
and Rivas-Saenz, 2009). The fields themselves are not a suitable
habitat for lizards and they occupy the adjoining walls and field
margins. During the entire field study only one juvenile lizard
was observed more than 1 m away from the nearest wall. The margins of the fields were of variable width (from 20 cm to 1 m) and
were characterized by the presence of spontaneous herbaceous
vegetation.
Podarcis bocagei (Seoane, 1884) is a small insectivorous lacertid
endemic to the northwestern Iberian Peninsula. It has a short lifespan (3–4 years) and occupies a variety of habitats, from wide open
forests with brushwood to coastal dunes and agricultural areas
(Galán, 1986). Adults show a strong sexual dimorphism; males attain larger sizes and a dorsal green coloration whereas females
have a brown pattern (Galán, 1986). They are active through most
of the year on sunny days, reducing their activity from mid November to February (Carretero et al., 2006). Juveniles hatch between
M.J. Amaral et al. / Chemosphere 87 (2012) 757–764
June and August and reach maturity in 1–2 years (Galán, 2004).
Individuals are considered adults if their snout-vent length (SVL)
is higher than 47.79 mm for males (Carretero et al., 2006) and
46.11 mm for females (Carretero, unpublished data).
2.2. Soil pesticide concentrations
Soil samples were collected for screening of a list of pesticides
and metabolites of present or past use and analyzed by LCMS–
MS (Liquid chromatography–tandem mass spectrometry). The
screening list included: (1) chemicals currently used in corn fields
in the area of study and some of their metabolites (glyphosate,
bentazone, dicamba, dimethenamid, alachlor, terbuthylazine, desethylterbuthylazine, mesotrione, chlorpyrifos, linuron); (2) products used in the recent past and their metabolites or which are
regarded as persistent (atrazine, desethylatrazine, heptachlor, lindane); (3) other chemicals of concern in the area because of their
intensive use, as signaled by the Direcção-Geral de Agricultura e
Desenvolvimento Rural (2,4-D).
Sampling was performed on two occasions in 2009, before and
after spring pesticide application by the farmers. Soil was sampled
every 50 m along the walls, on both sides, with at least four replicates per site. The samples were preserved in clean, new 200 mL
glass vials, covered with ParafilmÒ and sent for analysis (Marchwood Scientific Services, Southampton UK). The samples were air
dried and ground and a 2 g representative sample was extracted
with 20 mL of acetonitrile (containing 1% acetic acid). This was followed by a partitioning step with magnesium sulfate and a subsequent buffering step with sodium acetate. After mixing an aliquot
with methanol, the extract was injected into the LCMS–MS (Agilent
6410 Triple Quad LCMS–MS) without any clean up. Standards were
prepared in solvents at seven levels with recoveries in the range
70–120% (Level of Detection 0.5 lg).
2.3. Lizard sampling methods
Each of the localities were occupied by a subpopulation as defined by Wells and Richmond (1995). The six subpopulations were
sampled during two seasons (spring and autumn) over two consecutive years – 2008 and 2009. Sites were revisited over 3 d in each
season to ensure that a high proportion of the individuals were
sampled (2 d in the case of subpopulations Exp 2 and Exp 3 in
spring 2009). Sampling occurred during all day to ensure that captures were not biased towards either sex or juveniles, which tend
to be active at different times of the day (Galán, 2004). Some bias
towards the more conspicuous males was expected, but we believe
it was uniform across all sites and sampling times. At each visit,
individuals were captured by noose or hand, marked individually
by toe clipping and with non-toxic marker paint and released, noting gender, reproductive condition, snout-vent length (SVL, to the
nearest 0.1 mm), body mass (to the nearest 0.1 g on a Pesola scale)
and size class. Animals were visually inspected for tail break and
infection by ectoparasites. Digital photographs of the ventral side
were taken for analyzing asymmetries in scalation. Where there
was ambiguity in determining gender (i.e. subadults and juveniles), sex was confirmed by counting the number of transverse
rows of ventral scales; less than 28 in males, more than 28 in females (Kaliontzopoulou et al., 2005). Toe clipping has been considered the most appropriate method for marking small-sized lizards
as it is less stressful than the most frequently recommended technique to replace it – microchip implantation (Langkilde and Shine,
2006). Furthermore, toe loss is a common injury in natural subpopulations, and does not affect overall survival (Hudson, 1996). Care
was taken so that the longest toe on the hind foot was never amputated and that at least two other toes always remain intact on each
foot.
759
2.4. Data analysis
Population size and density (per 10 m) within each subpopulation was estimated in each sampling period assuming a closed
population scenario using the Schnabel or Petersen methodology
(Krebs, 1989). For the entire period of the mark-recapture study,
the Jolly–Seber open population model was employed (Krebs,
1989). Adult sex ratio was calculated for each subpopulation and
a Pearson’s Chi-square test determined departure from the expected 1:1 ratio.
For comparison of adult body size, estimated through the snoutvent length, SVL (see Meiri, 2010), two statistics were used: mean
body size, and the statistically more robust, mean asymptotic body
size, estimated as the 80th percentile (P80) (Brown et al., 1999).
Asymptotic body size is a preferred statistic for comparative
studies in which both sexes mature at small sizes because it is less
affected by proximate factors than mean size (Stamps and
Andrews, 1992; Brown et al., 1999). Following Roitberg and
Smirina (2006), we compared subpopulations, considering samples
of adult males/females collected from the same site in both years
and sampling seasons (2008 and 2009; spring and autumn). Mean
asymptotic body size was estimated by bootstrapping (100 samples for each sample size). Mean body size and mean asymptotic
body size were compared with a Kruskal–Wallis test, with mean
SVL/SVL P80 as the dependent variable and sampling site or subpopulation of origin as the independent factor. Body mass was
compared through an analysis of covariance (ANCOVA) with
sampling site as fixed factor and log SVL as the covariate. Tukey’s
multiple comparisons Post-hoc tests were employed to test for
significant differences when ANCOVA indicated significant
differences.
Fluctuating asymmetry (FA) in femoral pores has been widely
used with lacertids as a biomarker of environmental stress (López
et al., 2002; Crnobrnja-Isailovic et al., 2005). Femoral pores are
easy to identify and count and as such are advantageous over other
scalation parameters. Fluctuating asymmetry was estimated by triple independent counts of the number of femoral pores on both
sides of the body in adults and juveniles and subtracting the counts
of the right-hand side from counts of the left-hand side (R–L).
Counting was done using photos and always by the same observer
to avoid observer bias errors. Analysis of FA followed Palmer
(1994); data were inspected for bad raw measurements (defined
as errors of the operator during measurement or data introduction); abnormal individuals (individuals that have altered femoral
pores because of other conditions); and outliers. Outliers were
identified by the Grubb’s test statistic (Zar, 1996). Measurement
error (ME), the variability between replicate measurements, was
estimated and compared to FA by analysis of variance (ANOVA)
with sides and individuals as fixed factors. Test for antisymmetry
or directional asymmetry was not performed, as Podarcis lizards
exhibit ideal FA without evidence of any of those conditions (Lima
et al., 2008). Differences in absolute values of FA (|FA|) were tested
using Levene’s test for heterogeneity of variances among subpopulations (Palmer, 1994).
Parasite prevalence was estimated as the percentage of infected
lizards per size class and sex (males/females/subadults). Log-linear
models were used to analyze simultaneously the effects of site, sex,
and season on the presence of ectoparasites in adult lizards. We fitted different models to the data and analyzed the Pearson Chisquare statistics to choose the model that best describes the data.
All statistics were performed in STATISTICA (version 7.0, StatSoft Inc., USA) PC package (Hill and Lewicki, 2007). Assumptions
of homogeneity of variance and normality were examined, and
when data failed to meet these assumptions, transformations were
performed to fit the data to normal distributions or equal variances
where possible. Statistical significance was accepted at a = 0.05.
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3. Results
3.1. Soil pesticide concentrations
Only products that had been recently applied in the fields were
detected in our soil analyses, and all were herbicides, including the
active substances mesotrione, glyphosate, bentazone, dicamba,
dimethenamid-p, alachlor, and terbuthylazine (LOD = 0.5 lg,
Fig. 2, Table 1). The following substances were not detected in
our samples: desethylterbuthylazine, chlorpyrifos, linuron, atrazine, desethylatrazine, heptachlor, lindane and 2,4-D.
3.2. Population parameters
A total of 417 capture events (86 in site Exp 1, 40 in site Exp 2,
62 in site Exp 3, 73 in site Exp 4, 80 in site Ref 1 and 76 in site Ref 2)
including both captures and recaptures were recorded. Of these,
343 were unique individuals. Population size and density for each
of the subpopulations in each sampling time was estimated
through a closed model approach. No differences were found between exposed and reference subpopulations, but there was a seasonal difference. Subpopulation densities followed a general trend
of more animals in autumn than in spring (Fig. 3), with the exception of site Exp 4 in 2009. A reliable population size estimate for
the entire period of the study, which would account for changes related with stochastic events, was not achieved. Adult sex ratio
(deviations from 1:1), considering animals collected on both years
and sampling seasons, was significantly skewed in population Exp
4 and marginally non significant in subpopulations Exp 1 and Exp
3, with more males than females in the three subpopulations
(Pearson Chi-square tests, d.f. = 1: Exp 1 v2 = 2.9, p = 0.09; Exp 2
v2 = 0.4, p = 0.54; Exp 3 v2 = 3.5, p = 0.06; Exp 4 v2 = 6.0, p = 0.01;
Ref 1 v2 = 1.4, p = 0.23; Ref 2 v2 = 2.0, p = 0.15).
3.3. Body size measurements
Subpopulation Exp 2 was not considered in the analysis of body
measurements. This subpopulation had a low sample size (only 34
unique individuals), which impeded the calculation of an accurate
statistic for the characteristics under analysis. For male SVL, the
two statistics calculated (mean SVL and SVL P80) were highly correlated (r2 = 0.87, t-test = 0.84, p = 0.46). For both parameters we
detected a significant difference between subpopulations (Kruskal–Wallis mean SVL: H4,140 = 9.7, p < 0.05; SVL P80:
H4,500 = 331.3, p < 0.001) (Fig. 4). For mean SVL the post-hoc test
Fig. 2. Average pesticide content of soils (lg/kg soil): seven herbicides were
detected (LOD = 0.5 lg/kg) – Pre and post pesticide application in 2009 presented as
couplet bars (Exp – Exposed sites, Ref – Reference sites).
could not detect which subpopulations were contributing to the
difference, but a visual exploration of the data indicated Exp 3
and Ref 1. For P80 SVL, there were significant differences between
all sites. For females, mean SVL and P80 SVL were not so well correlated (r2 = 0.70, t-test = 2.57, p = 0.08). We could not detect any
significant difference between subpopulations for mean SVL (Kruskal–Wallis H4,91 = 2.9, p = 0.58) but a significant difference was detected for SVL P80 between all subpopulations, except between
Exp 1 and Ref 1 (Kruskal–Wallis H4,500 = 265.6, p < 0.001) (Fig. 4).
For male body mass, a significant difference was detected between the subpopulations (results not shown, ANCOVA Subpopulations: F4,114 = 2.8, p = 0.03). Despite the significance, the pos-hoc
test could not discriminate which subpopulations were contributing to the observed difference. Because of the high heterogeneity of
variances in female SVL we did not apply the test for female body
mass.
3.4. Fluctuating asymmetry
A visual inspection of the scatter plot of left vs right measurements revealed one outlier, which was not considered in the analysis. Between sides variation was significantly greater than
measurement error (ANOVA Side Individual: F1,329 = 14,
p < 0.001). The mean number of femoral pores varied between
16.3 and 17.8 and the absolute difference between sides between
0.6 and 0.8 (Table 2). Levene’s test for homogeneity of variances
did not show any difference between subpopulations or sexes
(F11,318 = 1.1, p = 0.35).
3.5. Ectoparasite prevalence
Ectoparasites were all acarids. Ectoparasite prevalence varied
between 27% and 63% in females, 7% and 40% in males and 23%
and 62% in subadults. Exp 4 lizards had the lowest infection rates.
A log linear analysis was performed to distinguish simultaneously
between the effects of site, sex and season of sampling on the prevalence of parasites in adults. The model that best described the
data indicated that the prevalence of ectoparasites was similar between subpopulations and seasons but that sex significantly affected their presence (Table 3). Females have generally a higher
prevalence of ectoparasites than males; in Ref 1 the prevalence
was similar.
4. Discussion
Studies that convincingly demonstrate whether environmental
contaminants have detrimental effects, or not, on natural reptile
subpopulations are still lacking. In agricultural areas in southern
Europe, lizards can be considered as valuable environmental indicator organisms, as they are conspicuous and have small home
ranges. Exposure to pesticides occurs mostly through ingestion of
contaminated food or soil, dermal exposure, and absorption by
eggs from surrounding environments.
All products detected on our sites were herbicides that had been
used in the past 2 years by farmers. Several other pesticides that
were used historically in the region and that would be expected
to persist were not detected. For example, atrazine, which was
intensively used in corn cultivation until the EU ban in 2007, and
its main metabolite (desethylatrazine) were not detected, despite
a relatively long residence time in soils (Huber, 1993). In response
to concerns for the environment and human health, modern pesticides have, as a generality, ‘evolved’ to degrade faster in the environment, be more specific in action, and thus, less toxic for non
target organisms. This woulpopulations to recover after their disappearance from soil. However, their continuous application in
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M.J. Amaral et al. / Chemosphere 87 (2012) 757–764
Table 1
Summary of chemical and ecotoxicological properties of the detected products. Degradation time (NP – non persistent, MP – moderately persistent) Classification of toxicity (?
possible; C – carcinogen, ED – endocrine disruptor, RDE – reproductive/developmental effects) (PPDB, 2009).
Alachlor
Bentazone
Dicamba
Dimethenamid-p
Mesotrione
Terbuthylazine
Glyphosate
Chemical class
Solubility (H2O 20 °C mg L
Chloroacetamide
Benzothiazinone
Benzoic acid
Chloroacetamide
Triketone
Triazine
Phosphonoglycine
240
570
250 000
1450
160
6.6
10 500
1
)
Fig. 3. Mean densities of Podarcis bocagei in each site. Each bar represents the mean
densities for two collecting seasons (2008 and 2009). Densities are based on either
Schnabel or Petersen N estimates (Krebs, 1989). Error bars represent standard
deviations (n = 2). Absence of an error bar for Exp 2-spring indicates that a density
estimate could not be calculated for 2009 (Exp – Exposed sites, Ref – Reference
sites).
Fig. 4. Mean SVL and SVL P80 of Podarcis bocagei adult females and males in the
different subpopulations, considering all sampling periods. Error bars represent
standard deviations (Exp – Exposed sites, Ref – Reference sites). Distinct letters
between sites indicate statistically significant differences (# P80 SVL p < 0.001; $
P80 SVL p < 0.001).
Volatility (25 °C mPa)
Degradation time (50 d)
Toxic classification
2.9
0.17
1.67
2.5
5.70 10
0.15
0.0131
14 (NP)
13 (NP)
8 (NP)
11 (NP)
5 (NP)
75.1 (MP)
12 (NP)
?C ?ED ?RDE
–
?RDE
?C ?RDE
–
?C ?RDE
–
3
agricultural sites and the interaction of the different products used
can put subpopulations at risk. As indicated by their presence prior
to re-application in 2009, alachlor, terbuthylazine and dimethenamid-p seem to be the most persistent of the chemicals applied,
and present themselves as unremitting chemical stressors for terrestrial fauna. Alachlor was detected in one of the reference sites
and the source was unknown; it is probably the result of some
cross-contamination through air from adjacent fields where pesticides were in use. No data are currently available for the effects of
any of these products on reptiles. Data available for birds have classified them as slightly to practically non toxic (PPDB, 2009); however, in a recent survey of literature, reptiles were more sensitive
than birds for more than 30% of the chemicals investigated (Weir
et al., 2010).
For vertebrate species, it is expected that agriculture intensification will have major effects on abundances within populations
rather than on overall biodiversity (Burel et al., 1998). Several studies have found higher levels of activity and density of individuals in
organic fields when compared to conventional systems (e.g. Hole
et al., 2005; Chamberlain et al., 2010). The majority of these studies
have found that the effects of pesticides are mostly indirect, acting
through changes in habitat quality related with different management practices. Although the use of pesticides is one of the characteristics of conventional agriculture, no direct correlation with
increases in diversity or abundance of mammals or birds was
found (Hole et al., 2005; Chamberlain et al., 2010). In the present
study, the activity pattern of lizard subpopulations was highly
irregular. Despite the high sampling effort, the overall rate of recapture was 15%, which made it difficult to estimate the overarching population sizes for the 2 year study period. Nonetheless, we
estimated population size and density for each sampling period,
which described an oscillatory trend in lizard abundance over
the seasons for each field. Lizard density was higher in autumn
than in spring because of juvenile recruitment. In subsequent
months high mortality rate among these juveniles have been reported (Galán, 1999, 2004). The subpopulation at Exp 4 was the
only population that did not follow this general trend; the autumn
population (as proportion of the spring population) was lower. This
was also the only subpopulation where the sex ratio was significantly different, with relatively few adult females captured in
spring of 2009. Male biased capture rates are generally expected
as males are usually more active and visible than females and
Table 2
Mean number of femoral pores in the right (FPR) and left (FPL) side of the body in the individuals of the six subpopulations. Abs. R–L – Absolute value of right minus left number of
femoral pores.
Site
n
Mean FPR
Mean FPL
Mean FP
Mean abs. (R–L)
Variance abs. (R–L)
Exp 1
Exp 2
Exp 3
Exp 4
Ref 1
Ref 2
74
33
42
61
58
62
16.65
17.83
17.13
17.28
16.58
16.32
16.87
17.89
16.63
17.39
16.55
16.33
16.76
17.86
16.88
17.34
16.56
16.33
0.84
0.61
0.74
0.74
0.75
0.79
0.45
0.35
0.37
0.41
0.61
0.46
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M.J. Amaral et al. / Chemosphere 87 (2012) 757–764
Table 3
Results of the log-linear subpopulation prevalence of ectoparasites in Podarcis bocagei by season (spring, autumn), age (adults, subadults) and sex (females and males).
Factor
d.f.
Partial Chi-sqr
p
Marginal associations Chi-sqr
p
Site
Season
Sex
Ectoparasites
Site season
Site sex
Site ectoparasites
Season sex
Season ectoparasites
Sex ectoparasites
Site season sex
Site season ectoparasites
Site sex ectoparasites
Season sex ectoparasites
Automatic model (site, ectoparasites)
5
1
1
1
5
5
5
1
1
1
5
5
5
1
11.828
0.686
3.744
19.819
7.446
3.178
5.503
2.113
0.295
4.063
4.600
3.086
1.750
1.463
Max. likel.
Pearson
Max. likel.
Pearson
0.037
0.407
0.053
<0.001
0.190
0.673
0.358
0.146
0.587
0.044
0.467
0.687
0.882
0.227
38.715
36.255
30.345
27.278
11.828
0.686
3.744
19.819
6.889
2.987
5.749
1.873
0.492
4.626
4.136
3.523
1.898
0.767
41
41
39
39
0.037
0.407
0.053
<0.001
0.229
0.702
0.331
0.171
0.483
0.031
0.530
0.620
0.863
0.381
0.573
0.681
0.838
0.921
Simplified model (sex-ectoparasites, site, ectoparasites)
the slightly skewed male:female ratio seen in all other sites does
not reflect the usual sex ratio in these kind of studies. It is possible
that the low number of females in 2009 and consequent low level
of juvenile recruitment in 2009 in Exp 4 occurred because of a stochastic event within the population, such as uneven predation, disease affecting only females or greater dispersal of juveniles.
However, Exp 4 is also the field with the highest pesticide load,
and we can speculate that the reduction in recruitment can be explained by mortality among females and/or lower recruitment in
juveniles as a direct or indirect consequence of this exposure.
Demographic studies have been suggested as appropriate tools
to assess the total effects of pesticide exposure. Even so, obtaining
accurate abundance estimates and relating them with pesticide
exposure is not an easy task, because slight variations in demographic parameters can influence population size. None of our
study sites represent closed systems, and re-colonization from
adjoining subpopulations is likely to be rapid if removal of individuals as a consequence of stochastic (e.g. a harsh winter or autumn)
or anthropogenic events (e.g. pesticide exposure) opens up new
opportunities for floating colonists (Amo et al., 2007). In our case,
most of the adjoining subpopulations occupy similar areas where
pesticides are in use; so all subpopulations are under the same
type of stress. Nevertheless, population demography in our study
cannot be considered as a very reliable or useful parameter to assess pesticide impact. Eventually long-term studies could overcome some of the problems we encountered.
In the absence of clear population indices based on simple
abundances of males, females and subadults, it is possible that
long-term exposure to pesticides will still be reflected in changes
in age structure. For example, lizards might not survive for so long
in exposed sites. In a study with the toad Bufo viridis, Sinsch et al.
(2007) found that age at maturity in males and longevity and postreproductive life span in females could be related with anthropogenic land use. Additionally, the hypothesis that older individuals
were being removed from the population seemed to be the most
probable explanation for a difference in population age structure
in bullfrogs inhabiting contaminated agricultural sites (Spear
et al., 2009). In a parallel skeletochronological study, we distinguished different age structures in the studied subpopulations
but we could not detect any clear relationship between herbicide
application and a differentiated age structure (Amaral, unpublished data). Alternatively, if energy needs to be re-allocated to
detoxification processes as a response to an increase of chemical
stressors in their environment, then this might be reflected in maximum body size attained during a lizard’s life. Generally, the subpopulations differed in terms of their body size measurements
for both females and males. Nevertheless, these differences did
not seem to be correlated with herbicide exposure. Changes in size,
can reflect differences in growth rates but factors that affect body
size have not been well studied. It is considered that body size is
affected by processes that differ from those affecting others aspects
of animal morphology (Brown et al., 1999). Factors known to influence body size are the size of the mother and conditions during egg
incubation (Braña and Ji, 2000). P. bocagei, like other oviparous lacertids, deposit their eggs in nests in soil. Egg deposition and incubation occurs during the period when pesticides are being applied
in the fields. Exchange of gasses and water with soil during egg
incubation can expose them to chemical contaminants. Marco
et al. (2004) found that fertilizers had a negative impact on hatchling body length in two Podarcis species. We did not study the
hatchling parameters in our subpopulations but we can hypothesize that they might be adding noise to our body size results.
Considering that no inbreeding processes are taking place in
this population (Pinho et al., 2007), fluctuating asymmetry is another morphological characteristic that can be related with developmental instability during the ontogeny of an individual and
ultimately be indicative of population stress (Sarre, 1996) and
proximity to pollution sources (Eeva et al., 2000). In our study,
no clear differences in FA in femoral pores were found between
the subpopulations. Our results are in accordance with another
study that found no differences in indices of FA in femoral pores
of different subpopulations of the lacertid Psammodromus
algirus exposed to a heavy metal spill (Márquez-Ferrando, 2008).
Other environmental factors, like egg incubation temperature
(Zhdanova and Zakharov, 2006) may have stronger effects on FA
than pollutants.
In the present study, we also could not detect any differences in
parasite presence between the subpopulations. From an ecological
perspective, parasite abundance and diversity could be reduced in
areas where pesticides are in use (Marcogliese et al., 2005). Alternatively, because exposure to some pesticides is suspected to affect
immune responses in some taxa, including amphibians (reviewed
in Mann et al., 2009) and mammals (Udoji et al., 2010), it can result
in higher infection rates (Carey et al., 1999). None of these scenarios were evident and the only difference observed was related with
sex, with females having generally higher prevalences of
ectoparasites.
In general, we detected few statistically significant differences
between the exposed and reference subpopulations. At least three
different scenarios can be proposed to explain the apparent absence of observable effects on population parameters. First, the
lack of differences may be because at the concentrations present
M.J. Amaral et al. / Chemosphere 87 (2012) 757–764
within the sites these pesticides have no observable effects upon
individuals. As indicated earlier, modern day pesticides are
increasingly more specific in action, thus reducing the likelihood
of effects on non target organisms. Eventually, if more potent pesticides, especially insecticides our results could be different. Second, individuals might have adapted to local stress conditions
and therefore survival can be related with acquired natural resistance to contaminants. Third, ultimately, other non toxicological
variables can mask pesticide effects.
To fully understand the impact of a contaminant under field
conditions it is important to consider also potential impacts of natural stressors. In some of our study sites, other lizard species of larger size that can act both as competitors (when juveniles) and
predators (when adults) were also present (Lacerta schreiberi and
Timon lepidus). Predation pressure was probably identical within
all sites as several snake and bird species were observed in the sites
or close vicinity. The presence of higher level predator species, in
particular snakes that have less mobility, might be indicative that
contaminant effects are not severe. Also, farming practices varied
slightly between fields, and these could be expected to influence
habitat quality. Reductions in habitat quality have been related
with changes in invertebrate communities and consequently in
food availability for vertebrates (Moreby and Southway, 1999).
Nevertheless, in a parallel study, we found that even if some of
the sites could be more favorable than others to Podarcis subpopulations, all fields supported arthropod communities sufficient to
sustain these subpopulations (Simão, 2011).
The detectability and activity levels of many reptiles, (highly
dependent on survey season), daily timing and environmental conditions have been pinpointed as reasons for the lack of field studies
in reptile ecotoxicology (Sánchez-Bayo, 2011). However, the main
reasons for this knowledge gap are probably related with the non
existence of protocols and model species for this group as there
are for other vertebrates. Moreover, the exposure and effects
parameters that have been used in the past with other species do
not seem the most adequate or applicable to lizard populations
in agricultural scenarios. For example, effects in reproductive
parameters impacted population dynamics in alligators exposed
to endocrine disruptor contaminants (Guillette et al., 2000). However, under agricultural field conditions these parameters are not
easily detected for lacertids, as they hide the eggs in soil inside
the walls at the margins of the fields. Reliable effect parameters
need to be found, especially for field situations.
Field studies should aim to detect correlations rather than effects, as they are not able to explain the complexity of field situations and can only tell us what is happening at a given time and
place. Nevertheless, the absence of observable effects on field subpopulations is indicative that lizard subpopulations are able to
cope or compensate for this level of exposure. In fact, P. bocagei
was the most widespread reptile species in this region, surviving
in the disturbed agricultural and peri-urban landscape while other
species disappeared (Ribeiro, 2011). It can be assumed that this
species is either inherently tolerant to the chemicals in their environment or have adapted to the stresses in the agricultural landscape over time. However, considering that there were
observable effects at the level of the individual animals, a set of
individuals from exposed and reference subpopulations were collected and brought to the lab for measurement of specific biomarker responses. The results of these tests are presented in a
companion paper (Amaral et al., 2012).
Acknowledgments
We appreciate the assistance of A.R. Agra, R. Ribeiro and CIBIO
members for help in animal collection. All animals were collected
under a permit issued by the Instituto da Conservação da Natureza
763
e Biodiversidade. This research was supported by FEDER through
COMPETE-Programa Operacional Factores de Competitividade,
and by National funding through FCT-Fundação para a Ciência e
Tecnologia, within the research project LAB-PET – Lacertid Lizards
As Bioindicators of Pesticide Exposure and Toxicity in intensive
market garden agriculture (FCT PTDC/AMB/64497/2006) and
through and FCT PhD Grant to M.J. Amaral (SFRH/BD/31470/2006).
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