Environmental Exposure to Polychlorinated Biphenyls and p,p´-DDE and Sperm Sex-Chromosome Disomy

Background: Chromosomal abnormalities contribute substantially to reproductive problems, but the role of environmental risk factors has received little attention. Objectives: We evaluated the association of polychlorinated biphenyl (PCB) and dichlorodiphenyldichloroethylene (p,p´-DDE) exposures with sperm sex-chromosome disomy. Methods: We conducted a cross-sectional study of 192 men from subfertile couples. We used multiprobe fluorescence in situ hybridization (FISH) for chromosomes X, Y, and 18 to determine XX, YY, XY, and total sex-chromosome disomy in sperm nuclei. Serum was analyzed for concentrations of 57 PCB congeners and p,p´-DDE. Poisson regression models were used to calculate incidence rate ratios (IRRs) for disomy by exposure quartiles, controlling for demographic characteristics and semen parameters. Results: The median percent disomy was 0.3 for XX and YY, 0.9 for XY, and 1.6 for total sex-chromosome disomy. We observed a significant trend of increasing IRRs for increasing quartiles of p,p´-DDE in XX, XY, and total sex-chromosome disomy, and a significant trend of increasing IRRs for increasing quartiles of PCBs for XY and total sex-chromosome disomy; however, there was a significant inverse association for XX disomy. Conclusions: Our findings suggest that exposure to p,p´-DDE may be associated with increased rates of XX, XY, and total sex-chromosome disomy, whereas exposure to PCBs may be associated with increased rates of YY, XY, and total sex-chromosome disomy. In addition, we observed an inverse association between increased exposure to PCBs and XX disomy. Further work is needed to confirm these findings.


Research
Chromosomal abnormalities are responsible for a substantial portion of reproductive problems. Unlike the standard indices of sperm quality, the consequence of successful fertilization with an aneuploid sperm are clear and serious, resulting in either congenital abnormalities or pregnancy loss. Sex-chromosome aneuploidies are the most common chromosomal error and are seen in at least 5% of clinical births, with some abnormalities not being realized until puberty (Hassold and Hunt 2001). Whereas maternal oocyte non disjunction accounts for > 95% of autosomal aneuploidy, sex chromosomal aneuploidy is known to be of paternal origin about 80% of the time (Hassold et al. 1996;Hassold and Hunt 2001). Sexchromosome aneuploidy is seen among otherwise normal mature sperm with a frequency between 0.13 and 1.20% (Egozcue et al. 1997); another study reported the average sexchromosome disomy across 23 studies to be 0.26% (Templado et al. 2005). Many studies have demonstrated an even higher frequency in infertile men (Bernardini et al. 1998;Calogero et al. 2001;Shi and Martin 2001). Although the rate and consequence of sperm sex-chromosome abnormalities are considerable, possible environ mental risk factors for sex-chromosome disomy have not been well investigated in epidemiologic studies (Wyrobek et al. 2005).
Much concern has been raised about organochlorine chemicals such as polychlorinated biphenyls (PCBs) and dichlorodiphenyl trichloro ethane (DDT) because of studies showing that both PCBs and dichlorodiphenyl dichloro ethylene (p,p´-DDE), the most stable daughter compound of DDT, are found in most of the population (Longnecker et al. 1997). Serum levels of PCBs and p,p´-DDE reflect long-term exposure. Depending on the congener type, PCBs have half-lives of 1-10 years, and p,p´-DDE has a half-life of ≥ 10 years.
PCBs and p,p´-DDE are persistent, lipophilic, endocrine-disrupting organochlorines that readily penetrate the blood-testis barrier (Bush et al. 1986). PCBs and p,p´-DDE could act via a number of possible mechanisms to alter endocrine homeo stasis, which is necessary for testicular function. For example, in vitro and animal models have demonstrated that PCBs can bind to estrogen receptors (Korach et al. 1988), increase gonadotropin-releasing hormone levels, or affect production and release of luteinizing hormone from the pituitary (Jansen et al. 1993). p,p´-DDE may affect spermato genesis through its anti androgenic activity (Kelce et al. 1995(Kelce et al. , 1997. In rodent models, changes in the endocrinologic environment of the testes have been shown to affect the rate of meiotic segregation errors (Mroz et al. 1999;Oppedisano et al. 2002). Furthermore, in epidemiologic studies, PCBs and p,p´-DDE have been associated with decreased semen quality, a key indicator of male reproductive health (Ayotte et al. 2001;Charlier and Foidart 2005;Dalvie et al. 2004;Hauser et al. 2003;Pant et al. 2004Pant et al. , 2007. Several other male reproductive toxi cants, such as non persistent pesticides (e.g., herbicides, fungicides, organophosphates) (Perry 2008) and benzene (Xing et al. 2010), have been associated with sperm aneuploidy, but no published studies to date have investigated the relationship between organochlorines and human sperm aneuploidy.
The present study was designed to investigate whether environmental exposure to PCBs and p,p´-DDE are associated with altered frequency of sperm sex-chromosome aneuploidy in adult men. We selected our study population without specific exposure (e.g., no known occupational exposure) to these compounds, because even small associations may have large public health relevance given the widespread exposures to PCBs and p,p´-DDE in the general population.

Materials and Methods
Subjects. Subjects were a subset of participants enrolled in a study assessing the impacts of environmental exposures on semen among 341 men. Men 20-54 years of age from couples seeking infertility evaluation from the Massachusetts General Hospital (MGH) Fertility Center between January 2000 and May 2003 were eligible for this parent study. Sixty-five percent of eligible men agreed to participate. Most men who declined participation cited lack of time on the day of the clinic visit as their reason for non participation. Men were excluded from the parent study if they were presenting for post vasectomy semen analy sis and/or receiving treatment for infertility (i.e., hormonal treatments). Each participant provided a semen and blood sample and completed a self-administered question naire that collected information on race/ethnicity, medical and fertility history, and lifestyle factors. Men were eligible for this sub study if they had an available semen sample for fluorescence in situ hybridization (FISH) analysis. Semen sample availability was based on whether an aliquot of sample was available for use from the bio repository, as this cohort has been used for other semen analysis research. There were no other additional eligibility requirements. Of the 341 men enrolled in the parent study, sufficient semen sample was available for 192 men (56%). The parent study was approved by the Harvard School of Public Health and MGH human subjects committees. All subjects signed an informed consent form prior to participation.
Semen analysis. All subjects were asked to abstain from ejaculation for 48 hr prior to giving their single semen sample by masturbation. Samples were liquefied at 37°C for 20 min before analysis. Analysis of fresh samples took place at the MGH Andrology Laboratory by andrologists blinded to exposure status. The volume, pH, color, and viscosity were determined for each fresh sample. Sperm counts and percent motility were first determined manually, then measured by computer-aided sperm analysis (CASA) using the Hamilton-Thorn Motility Analyzer (10HTM-IVO; Hamilton Thorne Inc., Beverly, MA). A minimum of 200 sperm from four different fields was analyzed. CASA provides data on rapid and linear motile sperm (percent), velocity of average path (micrometers per second), amplitude of lateral head displacement (micrometers), linearity, and other sperm movement characteristics. Additionally, each sample was prepared on two slides for morphological assessment using a Nikon microscope with an oil immersion 100× objective (Nikon Company, Tokyo, Japan). Sperm were scored as normal or abnormal using Kruger strict criteria (Kruger et al. 1988).
Serum PCB and p,p´-DDE measurements. On the same day as semen sample collection, non fasting blood samples were collected, frozen at-80°C, and subsequently analyzed by the Organic Chemistry Analytical Laboratory at the Harvard School of Public Health. The target analytes were 57 individual PCB congeners and p,p´-DDE. Details of the sampling, analytical, and quality control procedures are published elsewhere (Hauser et al. 2003;Korrick et al. 2000). Briefly, serum extracts were analyzed using gas chromatography with electron capture detection. Quantification was based on the response factors of individual PCB congeners and p,p´-DDE relative to internal standards. PCB concentrations were reported as individual congeners.
To adjust for variation in serum lipid levels among the nonfasting blood samples (Phillips et al. 1989), we measured percent lipid gravimetrically for each sample by weighing an aliquot of sample extract evaporated to dryness. Method detection limits (MDLs) were determined as recom mended by the U.S. Environmental Protection Agency (1984).
The MDL values for all PCB congeners and p,p´-DDE were < 0.05 ng/g, and most congeners had MDLs < 0.01 ng/g. X-Y-18 sperm FISH. One investigator (M.E.M.) performed FISH, imaging, and nuclei scoring analyses on frozen aliquots of all samples. She was blinded to exposure and covariate status while preparing and analyzing all samples. FISH was carried out for chromosomes X, Y, and 18 (autosomal control). Sex chromosomes were of primary interest, as they are most susceptible to disomy; however, 1818 disomy was also evaluated as a secondary outcome. Image processing and scoring were performed using custom software developed in MATLAB (Mathworks Inc., Natick, MA), using classification and scoring algorithms based on the scoring criteria described by Baumgartner et al. (1999). The overall hybridization efficiency was 97%, which is consistent with the hybridization efficiency reported by other groups (Johannisson et al. 2002;Martin et al. 1996;Tiido et al. 2005). These methods have been described in detail elsewhere (Perry et al. 2011).
Statistical analysis. Descriptive statistics for demographic characteristics and semen parameters were calculated. We used World Health Organization cutoffs (World Health Organization 1999) to categorize normal sperm concentration (≥ 20 million/mL), motility (≥ 50% motile), and morphology (≥ 4% normal). Because PCBs and p,p´-DDE were not normally distributed, nonparametric methods (Spearman correlation coefficients and Wilcoxon rank-sum tests) were used to explore their relationships with age, smoking status, abstinence time, race, and sperm concentration and motility.
We evaluated the relationship of sperm aneuploidy with p,p´-DDE exposure and with the sum of the four most prevalent individual PCB congeners (118, 138, 153, and 180; ∑ 4 PCBs). We also conducted exploratory analyses using other biologically relevant groupings of PCB congeners, as proposed by Wolff et al. (1997), considering those that are estrogen-like (congeners 44,49,52,101,187,174,177,and 157/201;∑estrogenic PCBs) and those that are dioxin-like (congeners 95/66,74,77/110,105/141,118,156,167,128,138,and 170;. We also considered multiple PCB groupings simultaneously, but it was difficult to distinguish their independent effects because of the high inter correlation among PCB subgroups. We used Poisson regression models to explore the associations between organochlorines and sperm sex-chromosome disomy. We fitted Poisson models using counts of each of the disomy types (XX18, YY18, XY18, total sex-chromosome disomy, and 1818) for each subject as the outcome variable, the natural logarithm of the number of sperm nuclei scored [log(nuclei)] as the offset variable, and organochlorine levels classified in quartiles as the exposure variable, adjusting for other relevant covariates. In Poisson regression, the offset variable allows for control of time/ space variation in the denominator, which in this study is the number of nuclei scored per subject. Because we found evidence of over dispersion (i.e., the variance exceeded the Poisson mean), robust methods based on generalized estimating equations were used to estimate exposure effects. Because of possible bias in standardizing exposure measures by serum lipids (Schisterman et al. 2005), we adjusted for serum lipids as a covariate in our models. Incidence rate ratios (IRRs) and 95% confidence intervals (CIs) were calculated using the first (lowest) quartile as the reference group.
We used scatter plots to explore the relation ships among covariates and PCBs and p,p´-DDE. Age, abstinence time, and smoking were included in all models based on a priori considerations about their likely associations with outcome or exposure (Blackwell and Zaneveld 1992;Hassan and Killick 2003;Vine 1996). Other potential confounding variables assessed were sperm concentration, motility, and morphology; race/ethnicity; and body mass index (BMI). Because both semen concentration and motility were highly skewed and residuals were not normally distributed, they were log-transformed prior to being included in models. Linearity after log transformation was verified with scatter plots. All potential confounding variables were included in initial models. Inclusion in final models was based on either a priori considerations, as mentioned above, or a change of at least 10% in the main exposure effect. We included p,p´-DDE in ∑ 4 PCB models, and vice-versa, to control for the other potential exposure.
Covariates in the final Poisson regression models were age (continuous), sperm concentration and motility (log-transformed continuous), abstinence time (≤ 2 days, 3-4 days, ≥ 5 days), serum lipids (continuous), and smoking status (never vs. ever). Additional models were fit to assess the appropriateness of assuming a linear relationship for the remaining continuous and ordinal covariates. Linear trend was determined by modeling quartiles of exposure as an ordinal variable assigned values of 1-4. Two-sided p-values < 0.05 were considered significant. We used SAS software, (version 9.2; SAS Institute Inc., Cary, NC) for all data analyses. Table 1 shows the demographic and semen parameter characteristics of the 192 study participants. Overall, participants had a mean (± SD) age of 35.4 ± 5.1 years and a mean (± SD) BMI of 28.1 ± 5.4 kg/m 2 . Most of the participants were white (85%), with 3% African American, and 4% Hispanic. Most men (75%) had never smoked, 15 (8%) were current smokers, and 33 (17%) were ex-smokers. Of the 192 men in our study, 21 (11%) had sperm concentrations < 20 million/mL, 84 men (44%) had < 50% motile sperm, and 35 men (18%) had < 4% normally shaped sperm. Participants in this study did not differ significantly with respect to their demographic or semen parameter characteristics compared with those in the full parent study cohort (n = 341) (data not shown).
T h e d i s t r i b u t i o n o f w e t -w e i g h t concentrations of p,p´-DDE and PCBs in serum and serum lipids is provided in Table 3. The median wet-weight concentrations of ∑ 4 PCBs and p,p´-DDE were 0.53 ng/g (range, 0.16-6.14 ng/g) and 0.97 ng/g (range, 0.23-30.3 ng/g), respectively. The median percent serum lipids was 0.5% (range, 0.13-1.17%). The median lipid-adjusted concentrations for ∑ 4 PCBs and were 110 ng/g lipids (range, 28.9-ng/g lipids) and 198 ng/g lipids (range, 34.8-7,776 ng/g lipids), respectively (data not shown). Serum levels of individual PCB congeners were lower than serum levels of p,p´-DDE (data not shown), a finding that is consistent with other U.S.based studies (Korrick et al. 2000). PCB and p,p´-DDE levels in the present study were comparable with the general U.S. population, as reported in the Fourth National Report on Human Exposure to Environmental Chemicals (Centers for Disease Control and Prevention 2009). Correlations among PCB measures, including the most prevalent individual congeners and congener groupings, were high (r = 0.6-0.9) (data not shown). However, correlations between p,p´-DDE and most PCB measures were more modest (r = 0.4).
After adjustment for age, smoking status, abstinence time, sperm concentration and motility, serum lipids, and ∑ 4 PCBs, serum p,p´-DDE levels above the lowest quartile were associated with increased rates of XX, XY, and total sex-chromosome disomy, but not YY disomy (Table 4). For example, the highest quartile of p,p´-DDE (≥ 1.68 ng/g) was associated with a 60% increase in the incidence rate of XX disomy (IRR = 1.6; 95% CI: 1.4, 1.7) compared with the lowest exposure quartile (≤ 0.61 ng/g). However, the dose response appeared non linear, with most of the increase in disomy occurring between the first and second quartile of serum p,p´-DDE and without substantial additional increases across subsequent quartiles. We observed no significant relationships between exposure to p,p´-DDE and 1818 disomy.
After adjustment for age, smoking status, abstinence time, sperm concentration and motility, serum lipids, and p,p´-DDE,    Includes congeners 6,8,16,18,25,26,28,31,33,37,41,44,47,49,52,60,95/66,70,74,77/110,84,87,97,99,101,105/141,118,128,135,136,138,146,149,151,153,156,157/210,167,170,171,174,177,180,183,187,189,194,195,196/203,199,206,209. b  we observed an overall significant linear increase in the rate of YY, XY, and total sexchromosome disomy for the second, third, and fourth quartiles of ∑ 4 PCBs compared with the lowest quartile based on a trend test (Table 5). However, not all disomy outcomes showed significant individual quartile increases. In contrast, there was a significant decrease in the rate of XX disomy above the first quartile of ∑ 4 PCBs. No significant relationships between exposure to PCBs and 1818 disomy were observed. In secondary analyses of alternative groupings of PCB congeners, after adjustment for age, smoking status, abstinence time, sperm concentration and motility, serum lipids, and p,p´-DDE, there was a significant increase in the rate of XY and total sex-chromosome disomy for the second, third, and fourth quartiles compared with the lowest quartile for ∑estrogenic PCBs and ∑dioxin-like PCBs and an increase in YY disomy for ∑estrogenic PCBs. However, we observed a significant decrease in XX disomy for higher quartiles of ∑estrogenic and ∑dioxin-like PCBs (data not shown).
We conducted sensitivity analyses for all disomy outcomes after excluding three men who had total sex-chromosome disomy rates that far exceeded current published ranges (total sex-chromosome disomy > 5%); these men were excluded to prevent undue statistical influence from these extreme values. In the reanalysis, the results remained essentially unchanged (data not shown). We also conducted a sensitivity analysis after excluding eight men with total nuclei scored < 1,000, as disomy estimates can be impacted by too few nuclei scored. Again, in the reanalysis, the results remained essentially unchanged (data not shown).

Discussion
This is the first epidemiologic study to investigate the relationship between environmental exposure to p,p´-DDE and PCBs with sperm sex-chromosome disomy. Our results suggest a significant increase in the rate of XX, XY, and total sex-chromosome disomy for the second, third, and fourth quartiles of of p,p´-DDE compared with the lowest quartile; these results persisted even after adjustment for potential confounders. For ∑ 4 PCBs, we observed a significant increase in the rate of YY, XY, and total sex-chromosome disomy after adjustment for potential confounders. In contrast, the rate of XX disomy was significantly decreased above the first quartile of ∑ 4 PCBs. We observed no significant relationships between PCBs or p,p´-DDE and 1818 disomy.
Most of the observed increase in disomy occurred between the first and second quartiles of serum PCB and p,p´-DDE exposures, with approximately constant increases across subsequent exposure quartiles. A saturation doseresponse relationship has been observed in other studies of organo chlorine health effects, including a recent study of organo chlorines and male peri pubertal growth (Burns et al. 2011). Furthermore, men in the first exposure quartile were not significantly different from men in higher exposure quartiles (e.g., with regard to demo graphic, lifestyle, or semen parameters; data not shown), so population differences across exposure groups are unlikely to explain the apparent threshold effect.
Because no prior studies have investigated organo chlorine exposure and sperm aneuploidy, comparisons of these findings with other studies are limited. Six epidemiologic studies have evaluated pesticide exposure and sperm aneuploidy (reviewed by Perry 2008); however, none have evaluated DDT or p,p´-DDE exposure, nor has an epidemiologic study evaluated PCB exposures and sperm aneuploidy. Of the four pesticide studies that investigated aneuploidy in the sex chromosomes (Recio et al. 2001;Smith et al. 2004;Xia et al. 2004Xia et al. , 2005, significant increased associations were seen with organo phosphate metabolites (Recio et al. 2001), ambient air fenvalerate levels (Xia et al. 2004), and ambient air carbaryl levels (Xia et al. 2005). These compounds are fast acting and have half-lives that range from days to weeks. One of the four pesticide studies did not find an association between insecticide, fungicide, or herbicide exposures and sex-chromosome diploidy (Smith et al. 2004). However, exposures in that study were determined by self-report and job exposure ratings and not from biomonitoring or environmental air monitoring.
Aneuploidy is most likely a consequence of disruption of meiosis during gametogenesis. It is unclear how hormone-disrupting  compounds such as PCBs and p,p´-DDE might interfere with the meiotic process. In studies of meiosis, Martin (2006) and Sun et al. (2008) have demon strated that infertile men have impaired chromosome synapsis, a significantly decreased frequency of recombi na tion, and an increased frequency of chromosomes that are missing a recombination site. Such errors make these cells susceptible to meiotic arrest and production of aneuploid gametes. Altered recombination has an effect on non disjunction (Ferguson et al. 2007;Shi and Martin 2001). Susceptibility of these chromosomes to nondisjunction may be due to reduced connections among homologous pairs in non recombinant chromosomes. A few toxicological studies have linked some pesticides to errors in somatic cell recombination (Giri et al. 2002;McCarroll et al. 2002), and chemicals known to disrupt hormone signaling may interfere with recombination sequences. In rodent models, changes in the endocrinologic environ ment of the testes have been shown to affect the rate of meiotic segregation errors (Mroz et al. 1999;Oppedisano et al. 2002). Alterations in endocrine homeo stasis induced by organochlorines could thereby adversely affect meiosis. However, identifying the likely mechanism(s) for an epidemiologic association between PCBs and p,p´-DDE with sex-chromosome disomy will require further investigation via experimental animal and in vitro studies. It is also unclear how increased exposure to PCBs might be protective for XX while increased exposure to p,p´-DDE appeared to increase XX disomy. Because sex-chromosome disomy is a rare event, it is susceptible to measure ment error, which could decrease the precision of the effect estimate. However, the error would be non differential and therefore not bias the results. Importantly, XY disomy was the most common disomy type and showed the most consistent relationship with exposures. XX and YY occurred less frequently than did XY. Less-consistent exposure associations with XX or YY may, in part, reflect even greater measurement error for these events.
Alternatively, in a study exploring the Y:X sperm ratio in men exposed to PCBs and p,p´-DDE, Tiido et al. (2006) showed an association between increased exposures to organochlorines and increased ejaculated Y-bearing sperm. Tiido et al. (2005) hypothesized that the decrease in X-bearing ejaculated sperm may be due to the effects of organo chlorines on the formation of X-chromosome micronuclei, which are chromosomal fragments or whole chromosomes that are not incorporated into daughter nuclei because of chromosomal breakage or dys function (Gauthier et al. 1999). A recent study by Pedersen et al. (2010) showed a positive association between dioxin-like PCBs and increased micro nuclei frequencies in cord blood of infants exposed in utero. PCB 118, a dioxin-like PCB congener that is a partial aryl hydrocarbon receptor agonist, was associated with an increase micro nuclei frequency in somatic cells of Belgian adults environmentally exposed to PCBs (De Coster et al. 2008). Although no studies have explored the impacts of micronucleation on meiotic cells (i.e., germ cells), increased X-chromosome micro nucleation, hypothesized to cause a decrease in X-bearing ejaculated sperm, might lead to decreased XX disomy. In addition, it is unclear whether exposure to PCBs versus p,p´-DDE and other organochlorines differentially impact micronuclei formation in germ cells, specifically sex chromosomes. Confirmation of these findings in other popu lations is needed.
The men in this study were members of sub fertile couples seeking infertility evaluation. Although they may differ from men in the general population, there is currently no evidence showing that they would differ in ways that would alter their response to PCBs or p,p´-DDE. Thus, our results may apply to general population samples as well. Participants were hetero geneous in their semen profiles, and more that half had normal semen parameters. The hetero geneity of men attending fertility centers should increase the generalizability findings from a clinicbased study population, but replication in other populations is needed to confirm the generalizability of our findings.
Chromosome disomy frequency estimates from the present study (mean ± SD: XX, 0.41 ± 0.40; YY, 0.38 ± 0.33; XY, 1.11 ± 0.83) were higher than previously reported ranges among men spanning the continuum of semen abnormalities using manual cell counting (XX, 0.03-0.37; YY, 0.04-0.21; XY, 0.06-0.42) (Egozcue et al. 1997;Templado et al. 2005). In a study of fertility clinic patients who were known recipro cal translocation carriers, XY disomy was as high as 4.1% (Moretti et al. 2009). Because we did not routinely perform karyo types, we do not know the prevalence of under lying reproductive conditions such as Robertsonian translocations in our clinic sample. Additionally, in a recent study of sperm disomy among andrology clinic patients using automated methods, Tempest et al. (2010) also reported higher total sex-chromosome disomy (averaging between 1.3 and 1.5% total disomy) compared with previously reported ranges (Egozcue et al. 1997). However, they found no differences between manual and automated estimates. Importantly, our previous validation studies comparing automated and manual results for sex-chromosome disomy in a small sample of normo zoospermic men from a fertility clinic also found higher disomy estimates than previously reported in men spanning the semen parameter continuum (Egozcue et al. 1997;Templado et al. 2005), but results did not differ between manual and semi automated methods (Perry et al. 2007(Perry et al. , 2011. Taken together, the use of a validated semi automated method for counting disomic sperm was considered a strength in that study because it allowed for objective processing of a large number of samples, which is important for etiologic hypothe sis testing.

Conclusion
The results of the present study suggest that men with higher serum p,p´-DDE levels have significant increases in the rates of XX, XY, and total sex-chromosome disomy. Men with higher serum levels of ∑ 4 PCBs had significant increases in the rates of YY, XY, and total sexchromosome disomy. In contrast, we observed a significant decrease in the rate of XX disomy for higher serum levels of ∑ 4 PCBs. It remains unclear why exposure associations differ among the various sex-chromosome disomy outcomes. This is the first epidemiologic study to investigate the relationship between environmental exposure to p,p´-DDE and PCBs and sperm sex chromosome disomy. Therefore, these findings require replication in other populations.