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research-article |
Department of Biology,3 Faculty of Science, Chulalongkorn University, Pathumwan, Bangkok 10330, Thailand
Department of Biology,4 Boston University, Boston, Massachusetts 02215
ABSTRACT
Contaminated groundwater plumes have formed on the Massachusetts Military Reservation (MMR), a Superfund site on Cape Cod, Massachusetts, as a result of chemical waste disposal. The plumes are of concern to the local people who rely on groundwater as a drinking water source. We used the freshwater turtle as a sentinel species to monitor the reproductive effects of exposure and, by inference, the potential for impact on human health. Our observations of male Chrysemys picta field-trapped from Moody Pond (an impacted site) and Washburn Pond (a reference site) on Cape Cod extended and supported prior observations of reproductive deficits. Morphometric comparison of precloacal length (PCL), which is a sexually dimorphic trait in the turtle, showed that Moody Pond males had a significantly longer PCL than Washburn Pond males. Moody Pond turtles showed reduced testicular weight, which was associated with significantly smaller seminiferous tubule diameter. Epididymal sperm counts were also markedly reduced in Moody Pond animals compared to Washburn Pond animals. Testicular histology and gonial proliferation, as determined by PCNA, were similar in both male populations, while the Moody Pond males had significantly higher germ cell apoptosis than the animals in Washburn Pond. These results suggest that a low-level mixture of xenobiotic contaminants impairs the reproductive functions of turtles exposed to the impacted site but not to the reference site environment.
environment, spermatogenesis, testis, xenobiotics
Global concerns have been raised over the adverse effects from exposure to xenobiotic substances that have the potential to interfere with the reproductive and endocrine systems. Examples of complex industrial contaminants that contain chemicals that affect the reproductive responses of wildlife populations have been reported in all major classes of vertebrates (reviewed in [1, 2]). Environmental contaminants may affect the endocrine and reproductive systems of organisms through several target sites. These chemicals can affect hormone synthesis, transport, metabolism, and activities in the gonad or components of the hypothalamic-pituitary-gonadal axis [2, 3]. Alternatively, contaminants can affect gonadal function through interference with germ cell progression and apoptosis during gametogenesis [4, 5]. Using wildlife as sentinel species not only provides information on the exposure and bioavailability of contaminants in the environment, but also provides important information on the effects of xenobiotics and the roles of other environmental factors in the final biologic response [6–8]. This also provides a critical link between the potential environmental risks to human health and reproductive success. The current study uses the male painted turtle (Chrysemys picta) as a sentinel species, to monitor potential environmental impacts on reproduction and development.
The area of concern is the Massachusetts Military Reservation (MMR), an 89-km2 active military facility on the upper western part of Cape Cod, Massachusetts (Fig. 1; [9]). This site was added to the US EPA national priority list as a designated Superfund site in 1989 as a result of organic and inorganic waste contamination at the base forming groundwater plumes to adjacent areas. To date, 80 potential contaminant source areas and 15 groundwater plumes flowing from the MMR have been identified [10, 11]. In the groundwater plumes, volatile organic compounds (VOCs), e.g., trichloroethene (TCE), perchloroethene (PCE), and ethylene dibromide (EDB), have been detected at levels above the state and/or federal maximum contaminant levels (MCL) allowed in the public drinking water supply [12]. Monitoring of surface water and sediments in the area has also revealed the presence of some pesticides and polychlorinated biphenyls (PCBs), as well as heavy metals at levels above the ecotoxicologic benchmark. Other organic and inorganic contaminants have been detected in the surface water and sediments albeit at levels below the ecotoxicologic benchmark, including PCBs (arochlor 254), semivolatile organic compounds (SVOCs), such as benzo(a) pyrene (BAP), and di(2-ethylhexyl) phthalate (DEHP) [13, 14].
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Previous observations of C. picta subpopulations near the MMR have suggested potential disruption of reproductive processes [15]. Experimental and analytical studies have suggested that low-level exposure to a xenobiotic mixture from the site or other source in the area may be involved. The present study was carried out to elucidate further the male reproductive and endocrine processes that are impaired by xenobiotic exposure. The parameters of interest included morphometry and gravimetry of the testes and accessory organs, levels of plasma testosterone, numbers of sperm, and spermatogenic parameters.
The study sites are in the vicinity of the MMR, a Superfund site on Cape Cod, MA. The potentially impacted site (Moody Pond; Fig. 1A) is a kettle hole pond with a maximum surface area of approximately 80 940 m2 and circumference of 960 m located near the southeastern border of the MMR and in proximity to the Eastern Briarwood and Fuel Spill-1 (FS-1) groundwater plumes. The Eastern Briarwood plume originates from an unknown source in the southern portion of the MMR, the most industrialized area of the MMR, and has been defined by TCE and EDB concentrations higher than the MCL. The FS-1 plume originates from the Aviation Gas Fuel Test Dump Site in the eastern part of the MMR and has EDB as the primary contaminant [11]. The reference site (Washburn Pond; Fig. 1B) is located east of the MMR, with a maximum surface area of approximately 28 330 m2 and circumference of 790 m, and is not in the path of the groundwater plume [13, 14]. These two study sites have similar ambient temperature, pH, and surrounding vegetation [16]. In the previous mark-recapture studies, we never found any animals that migrated between these two ponds. The Lincoln-Peterson estimates in 1998 gave a population of 187 ± 62 with a sex ratio of 1F:2.1M in Moody Pond, and 235 ± 53 with a sex ratio of 1F:2M in Washburn Pond (Rie and Callard, unpublished results).
Adult male turtles (Chrysemys picta) were collected from the impacted and reference sites during May–June 2002 and May–June 2003 (Massachusetts Division of Fisheries and Wildlife Permit for the collection of reptiles and amphibians no. 007.03SCRA) using sardine-baited circular mesh traps (Memphis Net and Twine, Memphis, TN) that were deployed for 48 h. Upon arrival in the laboratory, the animals were weighed and the following parameters were measured: straight carapace length (CL), plastron length (PL), straight carapace width (CW), height (H), and precloacal length (PCL). Turtles were kept at the AAALAC-accredited Laboratory Animal Care Facility (LACF) of the Department of Biology at Boston University. These animals were killed within 3 mo of capture (IACUC Assurance no. 98–022). Turtles were maintained under a 12L:12D cycle at 21°C with fresh running (uncontaminated) water and fed daily with turtle chow (Total Essential Nutrition Floating Food Sticks; Wardley Corp., Secaucus, NJ).
Tissue Collection and Processing
Turtles were anesthetized on ice and killed by decapitation. At autopsy, trunk blood samples were collected into heparinized tubes and centrifuged at 1000 x g for 20 min at 4°C to obtain plasma, which was frozen until assayed for testosterone. Testes and epididymes were collected and weighed. The testes from each individual were fixed in 10% neutral-buffered formalin, dehydrated, cleared, and embedded in Paraplast (Oxford Labware, St. Louis, MO). Gonad tissue blocks were sectioned at 10-µm thickness, mounted on poly-L-lysine-coated slides and used for histological (Mayer hematoxylin and eosin stain) and immunocytochemical examinations.
To measure sperm concentration, a small incision was made into the epididymis and 2 µl of the epididymal content was taken. The epididymal content was diluted 1:500 with sperm counting solution (0.6 M NaCl, 1% formalin, 0.5% crystal violet), loaded into a hemocytometer chamber, and counted under a light microscope. The total number of sperm counted in the hemocytometer chamber (volume of 0.1 µl) was converted to number per milliliter and multiplied by the dilution factor to give the sperm count per ml. The sperm count from each individual is given as the average of the values from the left and right epididymides. Sperm viability was examined according to the WHO protocol [17]. Two microliters of the epididymal fluid was diluted 1:50 with turtle Ringer solution (117 mM NaCl, 2 mM CaCl2, 3.9 mM KCl, 2 mM Na3PO4·12H2O [pH 7.12]). The content was stained with 200 µl of 1% eosin Y for 30 sec, followed by 300 µl of 10% nigrosin for 10 sec. Ten microliters of this mixture was placed on a glass slide, smeared, and air-dried. Three slides were prepared for each epididymis. Viability was assayed by counting 200 spermatozoa from each slide under the light microscope and differentiating the live (unstained/white) spermatozoa from the dead (pink-stained) spermatozoa.
Plasma Steroid Extraction and Radioimmunoassays
Steroids were extracted from plasma (120 µl) with 10 volumes of ethyl ether. After vortexing and clear separation of the two phases, the samples were snap-frozen in a dry-ice/ethanol bath. The upper ether layer was decanted into a clean assay tube and dried under a stream of air. The extraction steps were repeated and the steroid in the assay tube was reconstituted with 60 µl of RIA buffer (0.1 % gelatin in phosphate-buffered saline [PBS, pH 7.4]). Recovery of steroids was monitored by the addition of 10 000 cpm of tritiated tracer into duplicate aliquots of hormone-free plasma [18] prior to extraction, and counting the activity after reconstitution. The efficiency of each extraction, ranging from 69.08% to 92.28%, was used to correct the final concentration of steroid.
Serially diluted testosterone solutions or plasma extracts (25 µl) were incubated overnight at 4°C with 100 µl of tritiated tracer ([1,2,6,7-3H(N)]-testosterone; DuPont New England Nuclear, Boston, MA), 100 µl of anti-testosterone-11-BSA (obtained from Dr. G. Niswender, diluted 1:12 000 in RIA buffer), and 200 µl of RIA buffer. Free steroid was removed from each assay tube (except tubes used for total counts) by adding 400 µl of charcoal suspension (0.5% Norit A charcoal and 0.05% Dextran T70 in RIA buffer) and centrifuged at 1100 x g for 15 min. The supernatant was decanted into counting vials that contained 4 ml of BCS Scintillation Cocktail (Amersham Bioscience, Piscataway, NJ). After 20–30 min of equilibration, each vial was counted for 1 min in a scintillation counter (TM Analytic 6891). Once non-specific binding was subtracted, the percentage of bound tritiated steroid / maximum binding (B/B0) was calculated and plotted against the log concentration of testosterone (3.9–250 pg/25 µl), to generate a standard curve. Regression analysis was used to determine the testosterone concentration of the sample, with additional adjustments for recovery from the extraction and the dilution factor. The antibody directed against testosterone cross-reacts significantly with dihydrotestosterone (77%) and other androgens [19]. The sensitivity of the assay was 1.22 pg/25 µl. The coefficients of variation were 5.51%–7.66% for intra-assay variance and 13.02–13.06% for inter-assay variance.
Proliferating Cell Nuclear Antigen (PCNA) Immunocytochemistry
Testicular sections were deparaffinized, hydrated, and incubated with 3% H2O2 for 10 min. After rinsing in distilled water, the slides were pretreated with 1% zinc sulfate in a microwave oven for 6 min, to retrieve antigenic determinants. The sections were then washed in distilled water, followed by 0.05% Tween-20 in PBS (PBS-T). Nonspecific binding was blocked by adding 3% normal horse serum to each slide and incubating at room temperature for 30 min. PBS-T was used to wash the slides between each incubation step. Subsequently, the sections were incubated with mouse monoclonal antibody against PCNA (PC10; EMD Biosciences, La Jolla, CA) (1 µg/ml in PBS-T) overnight at 4°C. PCNA-positive cells were localized by the avidin-biotin-peroxidase complex technique (Mouse UniTect ABC kit; EMD Biosciences) using a 30-min incubation with biotinylated horse anti-mouse IgG and a 30-min incubation with ABC. Peroxidase activity was detected using 0.03% 3,3'-diaminobenzidine-4-hydrochloride (DAB) and 0.2% ammonium nickel sulfate in PBS. The slides were rinsed in distilled water, counterstained with hematoxylin, dehydrated, cleared, and mounted in DPX (distyrene and tricresyl phosphate in xylene; BDH Laboratory Supplies, Dorset, UK).
Apoptotic cells in the testes were detected by the TUNEL assay [20] using the TdT-FragEL DNA Fragmentation Detection Kit (EMD Biosciences). Sections were deparaffinized and hydrated with Tris-buffered saline (TBS, pH 7.6). The following incubation steps were carried out at room temperature unless mentioned otherwise, and TBS was used to wash the sections between each step. The sections were incubated with proteinase K (20 µg/ml in 10 mM Tris solution [pH 8.0]) for 20 min and with 3% H2O2 in methanol for 5 min. Subsequently, the sections were incubated in equilibration buffer (0.2 M sodium cacodylate, 0.03 M Tris, 0.75 mM CoCl2, 0.3 mg/ml BSA [pH 6.6]) for 20 min. Fragmented DNA was labeled by incubating sections with TdT enzyme and TdT reaction mixture at 37°C for 1.5 h. The reaction was terminated by submersing the sections in 0.5 M EDTA (pH 8.0). The sections were subsequently incubated with blocking buffer (4% BSA in PBS) for 10 min. Labeled DNA was detected by incubating the sections with peroxidase-streptavidin conjugate for 30 min. Peroxidase activity was visualized using 0.03% DAB. The slides were rinsed in distilled water, counterstained with 0.2% methyl green, dehydrated, cleared, and mounted in DPX.
Using PCNA and TUNEL as nuclear markers, the immunoreactive nuclei of spermatogonia were randomly counted under a light microscope. The gonial count was first recorded as the number of immunoreactive (IR) spermatogonia per seminiferous tubule based on an average of 50 tubules per section. The diameters of these randomly sampled tubules were measured using a micrometer. Finally, to compensate for the different sizes of the seminiferous tubules, the spermatogonial count is presented as the number of IR-spermatogonia per 10 000 µm2 of tubular area.
All the data were tested for normal distribution and homogeneity of variance. The Student t-test was used to determine significant differences from year to year before the data were combined into a single set. Morphometric data (CW, H, and PCL) were compared by analysis of covariance (ANCOVA) using carapace length as a covariable. Gravimetric data (gonad, epididymis) were compared by ANCOVA using body weight as a covariable. ANCOVA was followed by Bonferroni tests. Other site-related differences were compared using the Student t-test. A significant difference is reported at P < 0.05. The statistical analyses were carried out using SPSS version 10.0 for Windows.
Comparisons of body size by the Student t-test showed no significant differences in carapace length (CL; t = 1.69, df = 27, P = 0.10) or plastron length (PL; t = 1.68, df = 27, P = 0.10) between the populations. The analysis of covariance (ANCOVA) showed that the covariate, CL, was significantly related to carapace width (CW; F1,26 = 113.13, P < 0.05), height (H; F1,26 = 54.71, P < 0.05), and precloacal length (PCL; F1,26 = 37.73, P < 0.05). After controlling for the effect of carapace length, there was no significant difference between the populations in terms of carapace width (F1,26 = 0.05, P = 0.82) or height (F1,26 = 0.77, P = 0.39). However, male turtles from the impacted site showed significantly longer precloacal length (PCL) than males from the reference site (16.94 ± 0.75 mm in Moody Pond vs. 14.96 ± 0.40 mm in Washburn Pond; F1,26 = 5.04, P < 0.05) (Table 1).
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Gravimetric and Reproductive Parameters
There was no significant difference in body weights between the populations (227.71 ± 14.42 g in Washburn Pond vs. 226.50 ± 24.86 g in Moody Pond, t = 0.04, df = 21, P = 0.97). At autopsy in the Fall (September), a comparison of reproductive organ weight by ANCOVA showed that the body weight, a covariate, was significantly related to both testicular weight (F1,20 = 29.88, P < 0.05) and epididymal weight (F1,20 = 14.38, P < 0.05). After controlling for the influence of body weight, the ANCOVA showed a significant reduction in testicular weight for the Moody Pond males (0.38 ± 0.07 g) compared to the Washburn Pond males (0.59 ± 0.04 g; F1,20 = 6.30, P < 0.05). However, the epididymal weights were not significantly different between these populations (0.27 ± 0.02 g for Washburn Pond males vs. 0.21 ± 0.03 g for Moody Pond males; F1,20 = 3.06, P = 0.10). The epididymal sperm counts of the reference site animals were significantly higher than those of the impacted site animals (6.08 ± 0.53 x 109/ml vs. 3.53 ± 1.21 x 109/ml; t = 2.27, df = 21, P < 0.05). Sperm viability based on the percentage of live sperm was not significantly different between the populations (96.60 ± 0.65 % Washburn Pond males vs. 96.10 ± 1.05 % Moody Pond males; t = 0.41, P = 0.68). The radioimmunoassay for plasma testosterone showed no significant difference between the populations (0.86 ± 0.16 ng/ml in Washburn Pond vs. 0.95 ± 0.31 ng/ml in Moody Pond; t = –0.28, df = 21, P = 0.78) (Table 2).
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Comparison of seminiferous tubule diameter showed significantly smaller tubules in the Moody Pond animals (194.63 ± 4.58 µm) than in the Washburn Pond animals (231.44 ± 7.87 µm; t = 4.22, df = 21, P < 0.05, power of test [
= 0.05] = 0.96). Based on tubular development [21], spermatogenesis in the seminiferous tubules of these animals was at stage IV (spermiation), as indicated by release of sperm into the lumen of the tubule. Overall, there were no differences in spermatogenic stage between the populations.
Gonial Progression and Apoptosis
Proliferation of germ cells was detected by PCNA in every seminiferous tubule examined (Fig. 2A). Germ cells (primary spermatogonia) undergoing cell division were found near the myoid layer of the tubule. Comparison of PCNA-positive nuclei showed no difference between the populations (3.26 ± 0.70 nuclei per 10 000 µm2 tubular area in Washburn Pond animals vs. 2.35 ± 0.39 nuclei in Moody Pond animals; t = 0.93, df = 21, P = 0.39, power of test [ = 0.05] = 0.05). Apoptosis of germ cells was detected by the TUNEL assay near the basement membrane (Fig. 2B). Males from the impacted site showed significantly more apoptotic nuclei than males from the reference site (1.86 ± 0.44 nuclei per 10 000 µm2 tubular area vs. 0.69 ± 0.16 nuclei; t = –3.01, df = 21, P < 0.05, power of test [ = 0.05] = 0.76).
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In this study, multiple biologic endpoints were used to assess potential reproductive deficits in adult male C. picta, a sentinel species that was trapped from a potentially impacted pond near a Superfund site on Cape Cod, Massachusetts. Morphometric analysis was used to compare site-related differences in body size, an important life history trait that is often associated with other traits that directly affect reproductive success [22]. Based on body weight, carapace length, plastron length, carapace width and height, there were no significant differences in body size between the animals from these populations.
Precloacal length (PCL) is a sexually dimorphic trait in freshwater turtles and typically becomes longer in adult males than in adult females as the males reach maturity [23]. Comparison of PCL showed that the Moody Pond males had a significantly longer PCL than the Washburn Pond males, which suggests that this sexually dimorphic trait is sensitive to environmental factors. Although direct evidence is lacking, PCL seems to be hormonally determined and appears to be a sensitive indicator of xenobiotic exposure in turtles. In Chelydra serpentina, male turtles from areas contaminated with organochlorines (PCBs and pesticides) have shown reduced PCL compared to males from reference sites [24]. Additional studies of hatchling C. serpentina have shown that the precloacal lengths of hatchlings obtained from contaminated sites increase with body size at a slower rate than in hatchlings from a reference site, which suggests that the effect starts early in development and may result in permanent morphological change [25]. Although there were no differences in androgen levels between the animals from the two sites, an increase in PCL may be considered to be a response to an environmental contaminant. The active agents in this case have not been identified, although high cadmium levels prevail in the Moody Pond animals [26] and there is evidence that PCB-like contaminants are also present (see below). On-going investigations are focused on the chemical identification of xenobiotics of interest in sediments, water, and animal tissues from Moody Pond.
The reproductive parameters examined in this study suggest significant reproductive impairment in the impacted site population. Male turtles showed reduced testicular weight, which was corroborated by significantly smaller seminiferous tubule diameters. Epididymal sperm counts were also markedly reduced in the Moody Pond animals compared to the Washburn Pond animals. However, the weight of the epididymis, which is an accessory sex organ that reflects androgen status, was not different, in agreement with the similar levels of plasma testosterone in these populations. The levels of testosterone were comparable to the levels previously observed for C. picta during the Fall season [27]. The similarity of the testosterone levels between the populations confirms the results of previous studies [15] and suggests that the plasma testosterone levels are normal in male turtles at the impacted site.
Reproductive deficits in male turtles were further examined using markers of gonial proliferation and apoptosis, in order to establish whether the reduced sperm number is due to impairment of the spermatogenic process. The testicular histologies of turtles from both populations showed seminiferous tubules in the late stage of spermatogenesis, as expected for C. picta during the Fall season [21]. Gonial proliferation, as determined by PCNA, was similar in both male populations. In addition to cell proliferation activity, the number of maturing germ cells is influenced by the rate of apoptosis. The Moody Pond males had significantly higher germ cell apoptosis than the animals in Washburn Pond. In general, testicular apoptosis occurs naturally during spermatogenesis, possibly to reduce or limit the number of germ cells with genetic defects [28–30]. Apoptosis in the testis may also be enhanced by disruption of testicular function after exposure to environmental toxicants [30]. Since a lesion at the spermatogonial stages will subsequently lead to the depletion of more developed forms of germ cells [31], the enhanced gonial apoptosis found in the male turtles may lead to the reduction in sperm number seen for the Moody Pond animals. In addition, it is of importance to note the possibility of short-term effects that may not have been evident due to maintenance of animals in clean water in the laboratory.
Impairment of male reproductive functions may be caused by a direct toxicological effect on the testis or indirectly via the endocrine system [32]. Compounds that affect the hypothalamic-pituitary-gonadal axis may cause adverse effects on germ cells through Sertoli cells or directly on hormone receptors in or on germ cells [33]. The reduced testicular size and sperm concentration found in the present study are frequently associated with testicular toxicants [34]. One of the best known toxicants that directly targets the testis is cadmium [35], which is also a candidate toxin in the Moody Pond site. Based on a previous field study, adult animals from the impacted site had higher concentrations of cadmium in the liver, kidneys, and gonads, as well as higher levels of metallothionein-like protein, indicating exposure to heavy metal toxicants [26]. A study on the localization of isotopic cadmium in different tissues of adult C. picta 8 days after a single injection has shown a significant amount of cadmium in the reproductive endocrine tissues and reproductive tract, which implicates the reproductive organ as a potential target for cadmium toxicity in turtles [36]. Cadmium is also a recognized proapoptotic factor for the testis [30]. In vivo and in vitro studies using shark (Squalus acanthias) testes have shown that cadmium tends to accumulate mainly in testicular stem cells and spermatogonia and increases the rate of apoptosis in the testis [37]. In addition, cadmium is known to interact with specific amino acid residues and can exert xenoendocrine effects via the glucocorticoid receptor [38], estrogen receptor [39–41], and androgen receptor [42, 43]. Thus, cadmium may interfere with endocrine control of gonial progression and apoptosis in the gonad.
In addition to heavy metals, the observed reproductive endocrine interference in turtles from Moody Pond may be caused by direct or synergistic effects of organic xenobiotics in the area of the MMR. The chemical environment of the impacted site is poorly known, since there have not been any direct studies of this pond. However, several reports concerning areas near the MMR [11, 13, 14] suggest the presence of low-level mixtures of organic and inorganic xenobiotics at the site. Novillo and coworkers [44] have suggested the presence of PCB-like contaminants at this site, as determined by a change in aromatase transcription in a zebrafish bioassay. The preliminary chemical identification of xenobiotics of interest in turtle egg yolk by gas chromatography has revealed that the egg yolks from Moody Pond animals contain 91 µg/kg Dieldrin, which is a strongly suspected environmental endocrine disruptor [45], while the Dieldrin levels in yolks of Washburn Pond animals is below the detection limit of 34 µg/kg (Stavropoulos, Novillo and Callard, unpublished results). In a biomarker study, turtles from this site showed significantly higher activities of ethoxyresorufin-O-deethylase (EROD), glutathione-S-transferase (GST), as well as higher expression of cytochrome P4501A protein [16], indicating exposure to organic contaminants [46, 47]. Furthermore, organic contaminants may be metabolized to molecules that can cause cellular damage by binding to DNA or generating reactive oxygen species [1]. Oxidation of xenobiotics by CYP followed by conjugation with glutathione produces highly reactive ions as intermediates, which allow the formation of DNA adducts [48]. Thus, increased expression of CYP1A and GST in the impacted site animals [16] may cause reproductive toxicity due to metabolic activation of the parent compound.
The reproductive deficits observed in the present study corroborate and extend previous observations that suggest that low-level exposure to a xenobiotic mixture from the impacted site or other source in the area may disrupt the reproductive health of the turtle population, and may thus be of concern for humans who depend upon the water supply on Cape Cod.
ACKNOWLEDGMENTS
We thank Emily Marquez, Joseph St. George and Jim Sullivan (Boston University) for their assistance in the field.
FOOTNOTES
1Supported by a scholarship from Chulalongkorn University to N.K. and NIH grant ES 07381 to I.P.C. 
Correspondence: 2Noppadon Kitana, Department of Biology, Faculty of Science, Chulalongkorn University, Phyathai Road, Pathumwan, Bangkok 10330, Thailand. FAX: 66 2 218 5386; e-mail: nkitana{at}bu.edu
Received: 9 May 2006.
First decision: 11 June 2006.
Accepted: 19 October 2006.
REFERENCES
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