Thyroid disruption in the lizard Podarcis bocagei exposedto a mixture of herbicides: a field study

Rita C. Bicho • Maria Jose Amaral • Augusto M. R. Faustino •

Deborah M. Power • Alexandra Rema • Miguel A. Carretero •

Amadeu M. V. M. Soares • Reinier M. Mann

Accepted: 25 October 2012 / Published online: 10 November 2012

� Springer Science+Business Media New York 2012

Abstract Pesticide exposure has been related with thy-

roid disrupting effects in different vertebrate species.

However, very little is known about the effects of these

compounds in reptiles. In the Mediterranean area, lacertid

lizards are the most abundant vertebrate group in agro-

ecosystems, and have been identified as potential model

species for reptile ecotoxicology. The aim of this study was

to understand if the herbicides applied in corn fields have

thyroid disruptive effects in the lizard Podarcis bocagei.

Adult male lizards were captured in north-western Portugal

in corn fields treated with herbicides (exposed sites), and in

organic agricultural fields (reference sites). Thyroid and

male gonad morphology and functionality, and testosterone

levels were investigated through histological, immunohis-

tochemical and biochemical techniques. Lizards from

exposed locations displayed thyroid follicular lumens with

more reabsorption vacuoles and significantly larger follic-

ular area than those from reference fields. Furthermore,

testes of lizards from exposed locations had significantly

larger seminiferous tubule diameters, significantly higher

number of spermatogenic layers and displayed an up-

regulation of thyroid hormone receptors when compared

with lizards from reference areas. These findings strongly

suggest that the complex mixture of herbicides that lizards

are exposed to in agricultural areas have thyroid disrupting

effects which ultimately affect the male reproductive sys-

tem. Alachlor, which has demonstrated thyroid effects in

mammals, may be largely responsible for the observed

effects.

Keywords EDC � Thyroid gland � Testes � Reptile �Field study � Terrestrial ecotoxicology

Introduction

Endocrine disrupting chemicals (EDCs) include some

common environmental contaminants that can interfere

with the endocrine system and may affect normal devel-

opment, growth and sexual differentiation in a variety of

animal groups (Jugan et al. 2010; Diamanti-Kandarakis

et al. 2009). However, despite the increasing attention these

compounds are receiving, the impacts of EDCs on animal

populations and natural ecosystems remain poorly under-

stood (Diamanti-Kandarakis et al. 2009; Manning 2005).

EDCs can produce effects at doses even lower than those

used in the standard testing regimes of the U.S. EPA, or

those recommended by European Commission for the

Regulation of Chemicals (Vadenbergh 2004; Matthiessen

R. C. Bicho (&) � M. J. Amaral � A. M. V. M. Soares �R. M. Mann

CESAM & Departmento de Biologia, Universidade de Aveiro,

3810-193 Aveiro, Portugal

e-mail: ritabicho@ua.pt

M. J. Amaral � M. A. Carretero

CIBIO—Centro de Investigacao em Biodiversidade e Recursos

Geneticos, Universidade do Porto Campus Agrario de Vairao,

4485-661 Vairao, Portugal

A. M. R. Faustino � A. Rema

Departamento de Patologia e Imunologia Molecular, ICBAS,

Universidade do Porto, 4050-123 Porto, Portugal

D. M. Power

Centro de Ciencias do Mar (CCMAR), Campus de Gambelas,

Universidade do Algarve, 8005-139 Faro, Portugal

R. M. Mann

Centre for Environmental Sustainability, University

of Technology, Sydney, NSW 2007, Australia

123

Ecotoxicology (2013) 22:156–165

DOI 10.1007/s10646-012-1012-2

and Johnson 2007). Also, there are a multitude of physio-

logical processes that can be disrupted, which may only be

sensitive for relatively short periods during organogenesis.

Several EDCs are recognized for their effects on the

reproductive system, having estrogenic, androgenic, anti-

estrogenic and anti-androgenic effects on wildlife (Schmut-

zler et al. 2007; Jugan et al. 2010). More recently, the thyroid

axis was also recognized as a target for the action of endocrine

disruptors (Sciarrillo et al. 2008; Jugan et al. 2010; Schmutzler

et al. 2007). However, the complexity of the hypothalamus-

pituitary-thyroid axis and the numerous target organs for this

system means that there is a multiplicity of targets for the

action of chemicals with thyroid disruptive activities. As a

consequence, our understanding of the interaction between

environmental contaminants and the thyroid axis is still poor.

Most attention has been concentrated on the thyroid gland

itself, and it is now well established that some chemicals can

interfere directly with the synthesis of thyroid hormones and

its regulation by the hypothalamus and pituitary gland. Fur-

thermore, several histological biomarkers for increased or

decreased thyroid activity have been developed. For example,

increased follicular cell height and follicle area within thyroid

tissue are evident in thyroid glands that were actively stimu-

lated by the thyroid stimulating hormone (TSH) (Leatherland

1994; Hewitt et al. 2002).

Many pesticides are suspected EDCs and different lab-

oratory studies have demonstrated their effects on the

endocrine system of different taxa, including wildlife

populations heavily exposed to these contaminants

(Manning 2005; Meeker and Boas 2011). However, there

are limited data available that describe relationships

between pesticides exposure and altered thyroid function

(Meeker and Boas 2011). Among herbicides, some classes

such as the chloroacetanilide herbicides (e.g., alachlor);

triazines (e.g., atrazine); dinitroanalines (e.g., trifluralin)

and phenoxy acids (e.g., 2,4-D) have recognized thyroid

status-disrupting properties (Meeker and Boas 2011).

The disruption of thyroid gland homeostasis can also

affect the normal function or development of target organs

such as the testes (Zhang et al. 2009). Thyroid hormones

have an important role in regulation of testis development

and function in all life stages in mammals (Wagner et al.

2008; Jannini et al. 1995) and in reptiles (Haldar-Misra and

Thapliyal 1981; Plowman and Lynn 1973; Thapliyal et al.

1974), and this role is confirmed by the expression of

thyroid hormone receptors (TR) in testes (Cardone et al.

2000; Jannini et al. 1994; Buzzard et al. 2000).

Reptiles offer several useful attributes for discerning the

endocrine disrupting effects of pesticides and other chemicals

(Crain and Guillette Jr 1998). For example, reptiles appear to

have a lower capacity for detoxifying organic pollutants than

other vertebrates (Walker and Ronis 1989). However, com-

pared with other vertebrate classes, very little is known about

the effects of contaminants in these animals (Sparling et al.

2010). They are not routinely included in toxicity tests and

there is little information about their sensitivity to pesticides,

despite being identified by some authors as more vulnerable

than other taxa (Bruce et al. 2010; De Lange et al. 2009).

Moreover, most studies in reptile ecotoxicology concern

chemical tissue residues in field collected individuals

(Hopkins 2006; Linder et al. 2010). A far smaller proportion of

studies examined the effects of exposure to environmental

contaminants, and these are generally laboratory based

manipulations (McFarland et al. 2009; Capriglione et al.

2011). Although these studies are important to understand

responses, they do not provide enough information on the

effects on populations in complex natural environments

(Mann et al. 2009). Therefore, studies that examine the effects

of contaminant exposure in field populations are essential. In

Europe, lacertid lizards have been proposed as potential model

species for reptile ecotoxicology because of their high site

fidelity (Galan 1999; Carretero et al. 2006) and abundance in

agricultural areas (compared to mammals; Valverde, 1967)

and because they are relatively easy to maintain in captivity

(Mann et al. 2007; Cardone et al. 2008).

The current work forms part of a larger study on the

effects of corn pesticides in lacertid lizards. This larger

study included an examination of population indices and

biomarkers of pesticide exposure and effect among several

sub-populations of Podarcis bocagei (Amaral et al. 2012a,

b). The present study, aims to understand if pesticides

applied in agricultural fields have any thyroid disruptive

effects in the lizard Podarcis bocagei. We evaluated effects

on the thyroid gland and male gonad morphology and

functionality, and on the thyroid–gonadal axis function.

Materials and methods

Lizard collection and maintenance

Adult male P. bocagei lizards (n = 37) were collected in the

field margins of four agricultural areas in north-western

Portugal. The collection sites included two fields used for

corn growing where a mixture of pesticides has been applied

routinely for over 30 years (exposed), and two organic

agricultural fields with no history of pesticide application. A

detailed description of the collection sites: Exp 1, Exp 4,

Ref 1 and Ref 2 is published elsewhere (Amaral et al.

2012b). Lizards were collected from four sites to avoid

interference with population structures in any one popula-

tion. The major chemicals detected in the exposed sites were

alachlor, bentazone, dicamba, dimethenamid-p, mesotrione

and terbuthylazine (Table 1). Lizards were caught with a

noose during Autumn (November, 2009) corresponding to

the pre-hibernation period described for the species (Galan

Thyroid disruption in the lizard Podarcis bocagei 157

123

1995). Each animal was housed individually in glass terraria

(40 9 20 9 25 cm) in a climate control room. Each terrar-

ium contained a terracotta vase ([ = 16 cm) that provided

refuge and a basking location, and a 25 W incandescent lamp

(8 h/day) that created a thermal gradient (25–35 �C). Room

temperature was maintained at 22 ± 1 �C. Lighting in the

climate room was provided by external natural sunlight,

fluorescent lighting (2 9 40 W) and a high pressure sodium

lamp (400 W, 7000–12 000 Lx) for 12 h/day. Water was

provided in shallow dishes and renewed every 2–3 days, and

animals were fed with mealworms (Tenebrio molitor L.).

Lizards were held in captivity for 15 days during which

various parameters were measured (i.e., morphological,

behavioral and physiological); these data are described

elsewhere (Amaral et al. 2012a). Animal handling complied

with Portuguese animal ethics guidelines as stipulated by

Direccao Geral de Veterinaria and Instituto da Conservacao

da Natureza e Biodiversidade.

Dissection and histology

Lizards were killed by cervical dislocation. From each

animal, blood samples were taken and plasma was obtained

by centrifugation (800 rpm for 10 min at 4 �C) and stored

at -20 �C until hormonal analysis. Thyroid glands and the

left testis were removed, testes were weighed and tissues

were fixed in Davidson’s solution for 24 h. Tissues were

then washed in 70 % ethanol and stored in 70 % ethanol

until further processing. An automated tissue processor

(Microm STP 120) was used for paraffin embedding.

Sections (2 lm) were cut on a rotary microtome (Leica RM

2035). Thyroid sections were stained with Cleveland-Wolf

(CW) and testes sections stained with Hematoxylin and

Eosin (H&E) for light microscopic examination. Obser-

vations and photographs were made using an Olympus

BX51 microscope with an Olympus camera attached. To

perform thyroid histomorphometry, 20 follicles from each

animal were chosen randomly, and follicle area and follicle

cell height were measured. For testes histomorphometry,

20 seminiferous tubules from each animal were randomly

chosen for analysis. Spermatogenesis development was

assessed by measuring tubular diameter, and along this

axis, the width of Sertoli cells and the width of each

spermatogenesis layer (spermatogonia, spermatocytes I and

II, spermatids, spermatozoa). The number of spermato-

genesis layers observed in each animal was also recorded.

All measurements were carried out using ImageJ (Rasband

1997–2008).

Immunohistochemistry

Expression of thyroid receptors (TR) in testes was detected

by immunohistochemistry using the primary antibody—

TRa/b sc-772 (Santa Cruz Biotechnology, Inc. Santa Cruz,

Table 1 Average pesticide

content of soils (lg/kg

soil) ± SD: seven herbicides

were detected (LOD = 0.5 lg/

kg)—pre and post pesticide

application in 2009

Exp exposed sites, Ref reference

sites

Pesticide residue concentration

(lg/kg soil)

Sampled sites

Ref 1 Ref 2 Exp 1 Exp 4

Alachlor

Pre 10.23 ± 7.18 \0.5 7.03 ± 8.18 30.66 ± 8.25

Post 3.75 ± 3.75 \0.5 2.36 ± 4.13 16.97 ± 9.48

Bentazone

Pre \0.5 \0.5 \0.5 \0.5

Post \0.5 \0.5 39.88 ± 88.70 \0.5

Dicamba

Pre \0.5 \0.5 \0.5 \0.5

Post \0.5 \0.5 9.75 ± 27.75 \0.5

Dimethenamid

Pre \0.5 \0.5 13.52 ± 17.68 14.47 ± 12.26

Post \0.5 \0.5 0.59 ± 0.27 30.08 ± 28.13

Glyphosate

Pre \0.5 \0.5 \0.5 \0.5

Post \0.5 \0.5 \0.5 1.23 ± 1.55

Mesotrione

Pre 2.93 ± 4.85 3.58 ± 6.15 8.35 ± 9.71 \0.5

Post \0.5 \0.5 23.59 ± 34.12 20.7 ± 40.28

Terbuthylazine

Pre \0.5 \0.5 \0.5 \0.5

Post \0.5 \0.5 \0.5 33.48 ± 40.22

158 R. C. Bicho et al.

123

CA) as described in Virgilio et al. (2004). This is a rabbit

polyclonal antibody raised against a recombinant protein

corresponding to amino acids 1-408 that represents the full

length of TRa1 and reacts both with TRa and TRb iso-

forms. Testes sections were immersed in 1 L of extran

solution and microwaved for 20 min at 700 watts for

antigen retrieval. Antigen visualization was done with the

Novocastra Novolink Polymer Detection System (Leica

Microsystems GmbH, Wetzlar, Germany) and involved the

following steps: Sections were incubated with H2O2 (3 %)

for 10 min to eliminate endogenous peroxidase activity

followed by a 5 min incubation with a protein blocking

agent. Sections were subsequently incubated overnight at

4 �C with the primary antibody diluted at 1:45 with BSA

(5 %) and Triton X-100 (0.3 %), and on the following day,

washed in TBS-TRIS-buffered saline solution before

incubation for 30 min with the secondary antibody system

using diaminobenzidine (DAB) as a chromogen. Observa-

tions and photographs were made using an Olympus BX51

microscope with an Olympus camera attached. TR-positive

cells were identified as dark reaction product in the cell

nuclei, and their number was counted in 30 randomly

chosen seminiferous tubules.

Testosterone assays

Plasma testosterone (T) was determined using a specific

radioimmunoassay (RIA) as described by Scott et al.

(1984). Duplicate aliquots (50 lL) of diluted denatured

plasma (prepared as indicated above) were used in the T

assay. Preliminary analysis revealed parallelism between

dilutions of the samples and the standard curve. Intra-assay

and inter-assay coefficients of variation were 8 and 12 %,

respectively.

Statistical analysis

Quantitative data obtained from histological sections and radio-

immunoassays were tested for normality and homogeneity of

variance (Levene’s test). Differences between lizards from

exposed and reference sites (Type) were examined using analysis

of covariance (ANCOVA, a B 0.05) with log SVL as the

covariate. Differences in the number of spermatogenesis layers

were tested with the non-parametric Mann–Whitney test.

Results

Thyroid histology and histomorphometry

The thyroid gland of P. bocagei is a single discrete ribbon-

like structure that transversely crosses the middle of the

trachea, as described for other lizards. Thyroids of all

individuals presented a normal histological appearance—

colloid filled follicles with cuboidal epithelial cells sur-

rounded by interfollicular connective tissue and blood

vessels. However, in lizards from exposed sites, there were

a large number of reabsorption vacuoles in follicular col-

loid, whereas in lizards from reference sites these vacuoles

were scarce (Fig. 1a–c). Lizards from exposed areas also

displayed a significantly larger follicular area (Fig. 2a,

ANCOVA Type: F1,19 = 6.9, p = 0.02). No differences

were found in the follicular epithelial cell height (Fig. 2b,

ANCOVA Type: F1,19 = 1.5, p = 0.24).

Testis histology, histomorphometry

and immunohistochemistry

Macroscopically, testes from both reference and exposed

lizards had a normal appearance. Histological examination

of testes sections revealed that lizards from exposed sites

possessed a significantly higher number of spermatogenesis

layers (Figs. 3a,b, 4; Mann–Whitney test: z = -2.6,

p = 0.009). Specifically, eight lizards (&44 %) of this

group presented all spermatogenesis phases, from sper-

matogonia to spermatozoa; whereas, only one animal

(&6 %) from the reference locations presented all phases.

Moreover, individuals from exposed sites had a

Fig. 1 Thyroid gland sections of Podarcis bocagei stained with

Cleveland-Wolf. a Representative section of lizard from reference

site. b Representative section of lizard from exposed site with

enlarged follicular area, c High magnification (9400) showing

reabsorption vacuoles (arrow) in thyroid section of lizard from

exposed site

Thyroid disruption in the lizard Podarcis bocagei 159

123

significantly smaller layer of spermatogonia (Fig. 5a,

ANCOVA Type: F1,31 = 9.1, p = 0.005) and significantly

larger seminiferous tubule diameter (Fig. 5b, ANCOVA

Type: F1,31 = 4.5, p = 0.04).

Immunohistochemistry revealed expression and locali-

zation of TR proteins in both groups of lizards. In lizards

from reference locations, TR-positive staining was only

seen in Leydig cells and all spermatogonia and spermato-

cyte were TR-negative (Fig. 6a). In lizards from exposed

sites, TR-positive staining was seen in some spermatogo-

nia, spermatocyte and spermatids, and there was consistent

TR-positive staining in all mature spermatozoa (Fig. 6b).

Testosterone radioimmunoassay

No differences were found in the concentrations of circu-

lating testosterone between animals from exposed and

reference sites (Fig. 7, ANCOVA Type: F1,22 = 0.1,

p = 0.73).

Discussion

Thyroid disrupting effects related to exposure to herbicides

have been described in different species related to exposure

to herbicides (Mann et al. 2009; Meeker and Boas 2011;

Wilson et al. 1996). However, very little is known about

the effects of these compounds on reptiles (Cardone et al.

2008). Reptiles are an important non-target group of ver-

tebrate animals that will be exposed in regions of intensive

agriculture where many different pesticides are applied

annually, and in varying combinations. Although many

fauna persist in agricultural systems, it is unclear to what

extent they cope with the constantly changing cocktails of

these compounds. Prominent within the agricultural fields

of Portugal (and other Mediterranean countries) are abun-

dant populations of lacertid lizards. Their ubiquity in

Mediterranean agricultural ecosystems provides impetus

for the use of these species as biological models for eco-

toxicological studies in these environments.

An annual cycle in thyroid activity, influenced by cli-

matic variations and temperature, has been described for

several reptile species, including Italian Podarcis sicula, an

ecologically similar species to P. bocagei (Sciarrillo et al.

2000). In P. sicula, a peak in thyroid activity is observed

between March and July during the reproduction period,

and a minimum is recorded in winter before and during

hibernation (Sciarrillo et al. 2000). Our individuals were

collected in November, just before winter, and we expected

to find low thyroid activity. Indeed, individuals not

exposed to herbicides presented low thyroid activity.

However, lizards from exposed sites showed a more active

thyroid, with significantly larger follicular area, and a far

greater abundance of reabsorption vacuoles. Furthermore,

these lizards also displayed an up-regulation of TR in

testes. Similar findings were observed in amphibians from

agricultural areas, where individuals exposed to pesticides

presented increased levels of circulating T3 and T4

(Mosconi et al. 2005). Analogous histological and hor-

monal alterations were also observed in other animal

classes in different contamination scenarios. For example,

in mummichog (Fundulus heteroclitus), individuals from a

polluted habitat exposed to non essential metals and

organic materials had larger thyroid follicles, greater epi-

thelial cell heights, and elevated concentrations of circu-

lating T4 (Zhou et al. 1999).

The physiological mechanisms involved in the thyroid

effects of herbicides or other contaminants are not well

understood. Several pesticides that have known thyroid

disruptive activities generally decrease thyroid function

(Meeker and Boas 2011; Boas et al. 2006). Therefore,

within the context of the increased thyroid activity

observed in the present study, the chloroacetanilide herbi-

cides, a group that includes acetochlor, metolachlor and

Fig. 2 Thyroid gland histomorphometry results of Podarcis bocagei.a Follicular area of lizards from reference sites compared with lizards

from exposed sites. b Follicle epithelial cell height of lizards from

reference sites compared with lizards from exposed sites. *p \ 0.05;

error bars represent standard deviations, (n = 10 lizards from

reference sites; n = 12 lizards from exposed sites)

160 R. C. Bicho et al.

123

alachlor (among others), are of particular interest. In the

present study, lizards from contaminated sites were

exposed to a mixture of herbicides and alachlor was one of

the most persistent of the chemicals detected (Amaral et al.

2012b). In a study with rats, alachlor increased thyroid

activity (Wilson et al. 1996). In amphibians, acetochlor,

which is structurally very similar to alachlor, increased

thyroid activity (Helbing et al. 2006; Crump et al. 2002;

Cheek et al. 1999), and appears to be associated with

increased expression of TR (Veldhoen and Helbing 2001;

Helbing et al. 2006). It is hypothesized that the observed

alterations in thyroid function and the up-regulation of TR

in testes observed in lizards exposed to herbicides, were

consequent to alachlor exposure.

Reproduction in P. bocagei is characterized by an

annual spermatogenic cycle, which is influenced by climate

(Carretero et al. 2006). The maritime Atlantic climate of

north-western Portugal provides suitable temperature and

humidity conditions for P. bocagei activity during most of

Fig. 3 Cross sections of testes of Podarcis bocagei stained with

Hematoxylin and Eosin. a Representative section of lizard from

reference site: seminiferous tubules showing only 2 spermatogenesis

layers (spermatogonia (black arrow) and spermatocyte I (circled).

b Representative section of lizard from exposed site: seminiferous

tubule presenting all spermatogenesis layers (spermatogonia (blackarrow), spermatocyte I (circled) and spermatocyte II/spermatids (in

square) and spermatozoa (white arrow)

Fig. 4 Number of spermatogenic layers observed in seminiferous

tubules of testes of Podarcis bocagei. 2 layers include spermatogonias

and spermatocytes I; 3 layers include spermatogonias, spermatocytes

I and spermatocytes II/spermatids; 4 layers include spermatogonias,

spermatocytes I, spermatocytes II/spermatids and spermatozoa.

**p \ 0.01 Mann–Whitney test. (n = 16 lizards from reference sites,

n = 18 lizards from exposed sites)

Fig. 5 Testes histomorphometry results of Podarcis bocagei.a Width of each spermatogenic layer observed in seminiferous

tubule: A spermatogonia, B spermatocytes I, C spermatocytes

II/spermatids, D spermatozoa. b diameter of seminiferous tubules.

*p \ 0.05, **p \ 0.01; error bars represent standard deviations,

(n = 16 lizards from reference sites, n = 18 lizards from exposed

sites)

Thyroid disruption in the lizard Podarcis bocagei 161

123

the year including sunny days in winter (Carretero et al.

2006). As a consequence, the reproductive season starts in

February and lasts until July with spermatozoa production

starting between late November and December, peaking in

the middle of the breeding season and disappearing

towards its end (Carretero et al. 2006). Lizards in our study

were collected over two days in early November and were

subsequently maintained under identical conditions for

15 days before dissection.

Most lizards from reference sites presented only two

spermatogenesis layers (spermatogonias and spermatocytes

I) with a few presenting also the spermatocytes II/sper-

matids layer, and only one presenting all spermatogenesis

layers (including spermatozoa). In contrast, many of the

lizards from exposed sites had larger seminiferous tubules

and presented all spermatogenesis layers. These results

indicate that these lizards were in a more advanced stage of

spermatogenic activity than those from reference sites.

Furthermore, the spermatogonia layer was generally

smaller in exposed lizards, suggesting a higher recruitment

rate of these cells for further differentiation into more

mature germ cells when under active spermatogenesis. No

environmental factors offer any explanation for the

observed differences; temperature, relative humidity,

rainfall, and levels of solar radiation were all similar

between reference and exposed sites (Amaral et al. 2012b).

Therefore, it is possible that field exposure to the described

mixture of herbicides affected the normal spermatogenic

cycle in these lizards by triggering the beginning of sper-

matozoa production. The physiological mechanisms by

which the mixture of herbicides caused this effect in testes

are still unknown, and to the best of our knowledge, no

other studies have described inappropriate gonadal devel-

opment following exposure to environmental contami-

nants. However, some herbicides are able to affect

hormone metabolism and displace testosterone bound to

androgen receptors, ultimately changing normal testis

function and the progress of spermatogenesis (Cristino

et al. 2002; Cook et al. 1993; Cardone et al. 2008).

An annual cycle of sex steroids, including testosterone is

described for Podarcis sicula (Ando et al. 1992). Plasma

levels of testosterone reach a peak in the beginning of the

breeding season (March), starts decreasing towards the end

(July) and are low during hibernation period (Ando et al.

1992). In the present study, plasma titres of testosterone

revealed no significant differences between lizards from

different sites, suggesting that the spermatogenic activity

observed in lizards from exposed localities was probably

not induced through a mechanism involving increased

testosterone activity. However, several studies have dem-

onstrated both for reptiles and mammals that spermato-

genesis is under direct control of the thyroid gland (Jannini

et al. 1995; Holsberger and Cooke 2005; Cardone et al.

2008; Haldar-Misra and Thapliyal 1981). As indicted

above, in the present study, histology of the thyroid indi-

cated that this gland was significantly more active in lizards

from exposed sites, and that there was an up-regulation in

expression of TR in testes of these lizards. In a study with

P. sicula, the authors demonstrated that the induction of

Fig. 6 Expression and localization of thyroid receptor (TR) proteins

in testes of Podarcis bocagei. a Cross section of testis from reference

animal: positive nuclear immunostaining of Leydig cell (arrow).

b Cross section of testis from exposed animal: positive nuclear

immunostaining of spermatogonia (arrow), spermatocytes/spermatids

(circled) and spermatozoa (white arrow)

Fig. 7 Testosterone concentrations in Podarcis bocagei. Plasma

titres of testosterone (n = 10 lizards from reference sites, n = 15

lizards from exposed sites). Error bars represent standard deviations

162 R. C. Bicho et al.

123

androgen receptor (AR) mRNA was regulated by testos-

terone and T3 independently (Cardone et al. 2000).

Therefore, increased levels of T3 would increase the

expression of AR mRNA (Cardone et al. 2000), thereby

making sertoli and germ cells more responsive to circu-

lating testosterone. We can speculate that the premature

commencement of spermatogenesis in lizards exposed to

herbicides, which was not evident in lizards from reference

sites, was through the induction of AR mRNA, as a pre-

dictable consequence of higher thyroid gland activity and

up-regulation of TRs.

Amaral et al. (2012b) studied the demographics and mor-

phological aspects of natural field populations of P. bocagei

from these same agricultural sites. In general, few statistically

significant differences between exposed and reference sub-

populations were detected. However, among the same lizards

as were used in the present study, Amaral et al. (2012a)

examined various other physiological, behavioural and bio-

chemical biomarkers, and demonstrated that lizards exposed

to pesticides were affected by the exposure in a manner con-

sistent with an effect on the thyroid axis. Significantly

decreased body condition index, and marginally insignificant

increases in standard metabolic rate (SMR) were demon-

strated among exposed lizards (Amaral et al. 2012a). Avail-

able data has demonstrated that in lizards, SMR is influenced

by thyroid status (John-Alder 1984, 1990; Gupta and

Thapliyal 1985); and by sex hormones, as indicated by plasma

titres of testosterone (Gupta and Thapliyal 1985). Since in our

study no differences were observed in testosterone plasma

levels, our findings suggest that the increased SMR in exposed

lizards can be related to the disruption of the thyroid status

observed in these lizards. Furthermore, the lower body con-

dition index of exposed lizards could result from increased

energetic demands; an increase in SMR for maintenance

requirements has to be compensated for, either by an increase

in energy assimilation (i.e. feeding) or by a decrease in energy

available for growth, storage or reproduction (Amaral et al.

2012a; Carretero et al. 2006).

Conclusions

This study has demonstrated that histology of thyroid gland

and testes are valuable biomarkers for the evaluation of

endocrine disruption in lacertid lizards. Our results strongly

suggest that exposure of lizards to a complex mixture of

herbicides had thyroid disrupting effects that ultimately

affected the male reproductive system. Taking in consid-

eration that alachlor is known to have thyroid disruptive

effects in mammals, this herbicide may be largely

responsible for the observed alterations. It remains unclear

if the observable effects will affect the long term viability

of P. bocagei populations, or for that matter, other verte-

brate taxa living within the vicinity of agricultural fields.

Acknowledgments We appreciate the assistance of Ricardo Valente

and CIBIO members. All lizards were collected under a permit issued by

the Instituto da Conservacao da Natureza e Biodiversidade. This research

and the technical position of R.C. Bicho was supported by FEDER

through COMPETE-Programa Operacional Factores de Competitivid-

ade and National funding through FCT-Fundacao para a Ciencia 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). M. J. Amaral ben-

efited from a doctoral grant from FCT (SFRH/BD/31470/2006).

Conflict of interest The authors declare that they have no conflict

of interests.

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