Reproductive Toxicology 43 (2014) 1–7
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Exposure to polybrominated diphenyl ethers and male reproductive function in Greenland, Poland and Ukraine Gunnar Toft a,∗ , Virissa Lenters b , Roel Vermeulen b , Dick Heederik b , Cathrine Thomsen c , Georg Becher c,d , Aleksander Giwercman e , Davide Bizzaro f , Gian Carlo Manicardi g , k ´ Marcello Spanò h , Lars Rylander i , Henning S. Pedersen j , Paweł Strucinski , l m Valentyna Zviezdai , Jens Peter Bonde a
Danish Ramazzini Center, Department of Occupational Medicine, Aarhus University Hospital, Aarhus, Denmark Institute for Risk Assessment Sciences, Utrecht University, Utrecht, The Netherlands c Division of Environmental Medicine, Norwegian Institute of Public Health, Oslo, Norway d Department of Chemistry, University of Oslo, Oslo, Norway e Reproductive Medicine Centre, Skåne University Hospital, Lund University, Malmö, Sweden f Department of Life and Environmental Sciences, Polytechnic University of Marche, Ancona, Italy g Department of Life Science, Università di Modena e Reggio Emilia, Reggio Emilia, Italy h Laboratory of Toxicology, Unit of Radiation Biology and Human Health, ENEA Casaccia Research Center, Rome, Italy i Division of Occupational and Environmental Medicine, Lund University, Lund University Hospital, Lund, Sweden j Centre for Arctic Environmental Medicine, Nuuk, Greenland k Department of Toxicology and Risk Assessment, National Institute of Public Health–National Institute of Hygiene, Warsaw, Poland l Department of Social Medicine and Organization of Public Health, Kharkiv National Medical University, Kharkiv, Ukraine m Department of Occupational and Environmental Medicine, Bispebjerg University Hospital, Copenhagen, Denmark b
a r t i c l e
i n f o
Article history: Received 22 May 2013 Received in revised form 19 September 2013 Accepted 8 October 2013 Available online 26 October 2013 Keywords: Semen quality Reproductive hormones Inhibin B DNA damage Apoptosis PBDE BDE-47 BDE-153
a b s t r a c t Animal and a few human studies suggest that polybrominated diphenyl ethers (PBDEs) may affect male reproductive function. The aim of the present study was to evaluate if male reproductive function was associated with serum level of PBDEs. We evaluated, in a cross-sectional study, the effects of environmental exposure to BDE-47 and BDE-153 on reproductive hormones and semen quality, including markers of DNA damage and apoptosis, in 299 spouses of pregnant women from Greenland, Poland and Ukraine. Adjusted linear regression models indicated no strong associations between BDE-47 or BDE-153 exposure and markers of male semen quality or reproductive hormones. In the largest study to date we demonstrate that BDE-47 and BDE-153 exposure was not associated with altered semen characteristics or reproductive hormones, indicating that male reproductive function is not affected by the exposure level of these compounds in fertile European or Arctic populations. © 2013 Elsevier Inc. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs) have been used since the 1970s as additive flame retardants in a variety of consumer products including polyurethane foam used in mattresses and upholstered furniture and thermoplastics used in electronic
∗ Corresponding author at: Danish Ramazzini Center, Department of Occupational Medicine, Aarhus University Hospital, Norrebrogade 44, Building 2C, 8000 Aarhus C, Denmark. Tel.: +45 78 46 4251; fax: +45 7846 4260. E-mail address: [email protected] (G. Toft). 0890-6238/$ – see front matter © 2013 Elsevier Inc. All rights reserved. http://dx.doi.org/10.1016/j.reprotox.2013.10.002
equipment [1]. As they are mixed into polymers, and not chemically bound to the plastics or textiles, they are able to separate or leach from the products into the environment [2]. PBDEs are lipophilic compounds with a biological half-life up to 7 years [3]. Important sources of exposure include fatty fish and mammals from higher trophic levels, and therefore health risks for Arctic populations are a special concern [4]. House dust has also been indicated as an important contributor to human PBDE exposure, especially for toddlers and small children [5,6]. Of the 209 PBDE congeners, BDE47 is usually the compound detected in the highest concentration in human biological material, followed by BDE-153 [5]. In addition to effects on the thyroid hormones [7], several PBDEs have shown the
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potential to interfere with reproductive hormone signaling through anti-androgenic, estrogenic or anti-estrogenic activity in in vitro models [8–10]. In vivo subacute (28 day) toxicity tests on adult male rats exposed to a commercial pentaBDE mixture in a dose range from 0.27 to 200 mg/kg body weight/day have shown dosedependent decreased weight of the epididymis, seminal vesicles and prostate, and an increase in sperm head deformities with lower confidence bound of the estimated benchmark dose in the range of 10–50 mg/kg body weight/day [11]. Exposure of the fish Fathead Minnows (Pimephales promelas) to 14 g BDE-47/day for 25 days indicated specific effects on male reproductive function, including a reduction in sperm concentration [12]. Furthermore, exposure in ovo of American kestrels (Falco sparverius) to the commercial pentaBDE mixture, DE-71, measured at 1 g/g wet weight in the eggs, reduced testosterone levels and altered testis structure in the male offspring [13]. Some earlier toxicological evaluations of PBDEs did not indicate adverse effects on reproductive function in rodents after exposure in adulthood [14], but more recent experiments in mice have reported that, whereas neonatal exposure to 500 or 1500 mg/kg BDE-209 per day could induce weak effects on a few aspects of sperm function [15], in utero exposure to BDE209 induced testicular changes together with a significant increase of sperm head abnormalities and sperm chromatin damage in the male progeny [16], and some of these effects were observed at a much lower concentration (10 mg/kg/day BDE-209). In humans, only a few studies addressed the issue of environmental PBDE exposure in relation to adult male reproductive function. A Japanese study of 10 young men indicated adverse associations between BDE-153 and sperm concentration and testis size [17], and a Canadian study of 52 men recruited in an infertility clinic indicated adverse associations of BDE-47, BDE-100 and the sum of measured PBDE congeners (BDE-47, BDE-99, BDE-100 and BDE-153) on sperm motility [18], but no relation with other semen parameters. Finally, an American study of 62 men indicated positive associations between house dust penta-BDEs and serum level of estradiol and sex hormone binding globulin (SHBG), and inverse associations with follicle stimulating hormone (FSH). In the same study, house dust octa-BDEs were positively associated with luteinizing hormone (LH) and testosterone and finally deca-BDEs were inversely associated with testosterone [19]. This study was expanding on a previous report from the same group indicating an inverse associations between the measured PBDE congeners (47, 99 and 100) in dust samples and free testosterone, LH and FSH [20]. Thus, although epidemiological studies are only partly supported by experimental studies, animal and human studies suggest that PBDEs may affect human reproductive health. The associations have so far only been evaluated in small studies with divergent endpoints, and confirmation is needed in larger studies before more definite conclusions on effects on male reproductive health can be made. The aim of the present study was to evaluate if PBDE exposure is associated with male reproductive function in three populations of fertile men with considerable variation in PBDE exposure.
2. Materials and methods 2.1. Study populations The present study is part of a European study on fertility, the CLEAR study (www.inuendo.dk/clear), using a uniform protocol for data collection in Greenland, Kharkiv in Ukraine, and Warsaw in Poland [21]. Six hundred and two partners of pregnant women provided a semen sample and a blood sample, and filled in a questionnaire on lifestyle, occupation and medical history. All men were 18 years
or older at the time of enrolment. In each of the three countries 100 blood samples were selected randomly among the samples with sufficient volume to allow analysis (>4 mL). One sample was lost due to problems during sample preparation before chemical analyses, and therefore the final sample size used to evaluate the effect of lipid adjusted PBDEs on male reproductive outcomes was 299. The 299 included men were not significantly different from the 303 non-included men regarding measures of semen quality or reproductive hormones. The participation rates in the study varied from 29% in Warsaw and 33% in Kharkiv to 79% in Greenland. Study populations and data collection procedures have been described in detail elsewhere [21]. 2.2. Collection of semen samples and basic semen analysis All semen samples were collected by masturbation. The subject was asked to abstain from sexual activities for at least two days before collecting the sample, and to report the actual abstinence time. The sample was kept close to the body to maintain a temperature close to 37 ◦ C if transport to the local hospital after collection was necessary. Analysis of semen samples was initiated within one hour after ejaculation for 83% of the samples. The samples were analyzed for concentration, motility and morphology according to a manual for the project based on the 1999 World Health Organization (WHO) manual for basic semen analysis [22], which was the most recent version of the manual at the time of sample collection. Briefly, sperm concentration was determined in duplicate using an Improved Neubauer Hemacytometer. Sperm motility was determined by counting the proportion of (a) rapid progressive spermatozoa; (b) slow progressive spermatozoa versus (c) nonprogressive motile spermatozoa; and (d) immotile spermatozoa among 100 spermatozoa within each of two fresh drops of semen. Semen samples from Warsaw and Kharkiv were analyzed for concentration and motility at one central hospital in each region, whereas the samples from Greenland were analyzed at the local hospitals or nursing stations spread across the country. One person performed all the semen analyses in each of the three countries. These three persons were previously trained in a quality control program set up specifically for this study [23]. The inter-observer variation as consider concentration was found to be 8% and for motility 11%. Spermatozoa from all populations were stained and analyzed for morphology centrally at Skåne University Hospital in Malmö, Sweden. Abnormalities were classified as head defects, midpiece defects, tail defects, cytoplasma drop and immature spermatozoa. The morphology was evaluated for at least 200 sperms in each sample by two technicians taking part in the NAFA-ESHRE (Nordic Association for Andrology – European Society of Human Reproduction and Embryology) external quality control program. 2.3. Sperm chromatin structure assay (SCSA) DNA damage was measured by SCSA following the standardized procedure described by Evenson et al. [24]. Briefly, all coded frozen semen samples from the three study regions were shipped on dry ice to the flow cytometry facility of the ENEA Laboratory of Toxicology (Rome, Italy) for SCSA analysis; a flow cytometric (FCM) technique which identifies the spermatozoa with abnormal chromatin packaging, envisioned by increased susceptibility to acid-induced DNA denaturation in situ followed by acridine orange fluorescence staining [24,25]. Adhering strictly to the SCSA protocol as described by Evenson et al. [24], the DNA fragmentation index (%DFI) – representing the percentage of sperm with detectable DNA breaks – together with the percentage of sperm with high levels of high DNA stainability (%HDS) – representing the fraction of sperm with anomalies in the histone-to-protamine transition – were
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calculated by a dedicated software (SCSASoft; SCSA Diagnostics, Volga, SD, USA). FCM measurements were stopped when a total of 10,000 spermatozoa had been accumulated for each sample. For the flow cytometer set-up and calibration, a reference semen sample retrieved from our laboratory repository was used as described in [25]. Inter-day SCSA variability for %DFI, based on 216 flow sessions showed a coefficients of variation (CV) of 6.0%, whereas the CV for the %HDS was 4.8%. In addition, 358 randomly chosen samples (50.6% of the total) were measured twice in independent FCM sessions. Results from the two measurements were highly correlated (DFI, r = 0.96). External quality control indicated a median inter-sample variability for %DFI, expressed as CV, of 1.5% [25]. 2.4. Determination of DNA fragmentation by terminal deoxynucleotidyl transferase dUTP nick end labeling (TUNEL) assay
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0.2 IU/L, 0.2 IU/L, and 8.0 pmol/L, respectively. The total assay coefficients of variation were 2.9%, 2.6%, and 8.1%, respectively. Serum testosterone levels were measured by means of a competitive immunoassay (Access; Beckman Coulter Inc., Fullerton, CA, USA) with an LOD of 0.35 nmol/L and a total assay CV of 2.8% at 2.9 nmol/L and 3.2% at 8.1 nmol/L. SHBG concentrations were measured using a fluoro-immunoassay (Immulite 2000; Diagnostic Products Corporation, Los Angeles, CA, USA). The LOD was 0.02 nmol/L. The total assay coefficients of variation were 5.5% and 4.6%, respectively. Inhibin B levels were assessed using a specific immunometric method, with a detection limit of 15 ng/L and intraassay and total assay coefficients of variation 40%), so the median blank contamination level of 1.58 pg/g serum was subtracted from all individual measurements. BFR concentrations were analyzed as lipid adjusted values. Serum concentrations of triglycerides and cholesterol were determined by enzymatic methods [30]. The total lipid concentration in serum (g/L) was calculated as
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total = 0.96 + 1.28 × (triglycerides + cholesterol), as described in [31]. 2.9. Statistical analysis We evaluated trends of exposure-outcome associations by general linear regression models across the three regions. Only exposures with measurements above the LOQ in more than 70% of the study population samples were included in statistical analyses [32], limiting the analyses to BDE-47 and BDE 153, detected above LOQ in 99% and 97% of samples, respectively. Values for samples below LOQ for these compounds were estimated and included in the statistical analysis. The values were estimated using iterative single imputation, in which the estimated means depended on the other measured PBDE exposures and study population to allow for population-specific residual variances [32]. Due to the skewed exposure distributions, we transformed PBDEs by taking the natural logarithm (ln). To improve normality and homogeneity of variance of the residuals in the tested associations, all outcome variables were ln-transformed. Model fit was evaluated by visual inspection of residuals, i.e. observed versus predicted values. Semen variables (sperm concentration, total sperm count, volume, motility, morphology, DFI, HDS, TUNEL, FAS and BclxL) were, according to an a priori decision, adjusted for potential confounding effects of known determinants of male reproductive function, including for all semen variables: abstinence time and age (both ln-transformed); current smoking (yes/no); ever urogenital infections (yes/no); body mass index (BMI – kg/m2 ) (25), and country [21,33–36]. Sperm concentration, motility and morphology were additionally adjusted for spillage (yes/no). Furthermore, analysis of sperm motility was restricted to samples analyzed within one hour of collection (n = 250), and volume and total sperm count were restricted to samples from individuals reporting no spillage (n = 267). Associations between PBDEs and reproductive hormones were adjusted for age, smoking, BMI, country and blood sampled before noon (12:00) or after. Effect modification by country was tested by interaction tests, by including a country × exposure term in the regression models including covariates as specified in the adjusted regression models for each exposure-outcome association tested. Exposure-outcome associations were additionally evaluated stratified by country in a supplementary analysis. In addition, supplementary analyses were run without country in the models. Statistical analyses were performed using the SAS statistical software version 9.1 (SAS Institute Inc., Cary, NC, USA). 3. Results 3.1. Characteristics of the study populations In total 299 men delivered a semen sample and had a blood sample analyzed for PBDEs. The median PBDE concentrations were up to 10 fold higher in serum samples from men from Greenland compared to Ukraine, with intermediate levels of the measured compounds in Poland. Also, other characteristics like age, percentage of persons reporting spillage, percentage reporting previous urogenital infections and percentage of the populations smoking varied considerably across countries (Table 1). 3.2. Associations between PBDE exposure and male reproductive function Markers of male reproductive function in relation to BDE-47 and BDE-153 are shown in Tables 2 and 3. In crude models, the measured PBDE congeners were inversely associated with estradiol, testosterone, FAI, and one marker of sperm DNA damage (DFI). In
addition BDE-47 was inversely associated with inhibin B, and BDE153 was positively associated with FSH, SHBG and DNA damage measured by the TUNEL assay. After adjustment for age, abstinence time, smoking, urogenital infections, BMI and country, none of the negative associations between BDE-47 or BDE-153 exposure and male reproductive health outcomes remained. However, we observed an inverse association between BDE-47 exposure and HDS after adjustment ˇ = −0.05 CI (−0.09; −0.01). No significant effect modification by country was indicated. Results stratified by country are shown in Supplementary Tables 2 and 3. Also in the stratified analyses, an inverse association between BDE-47 and HDS was found in all countries, although only statistically significant in Ukraine. Apart from this no consistent associations were found across countries. However, a few significant associations in single countries were observed, including for BDE-47: an inverse association with semen volume in Greenland but a positive association with semen volume in Poland, and a positive association with TUNEL in Poland and a positive association with Bcl-xL in Ukraine (Supplementary Table 2). For BDE-153, a positive association with FSH was observed in Ukraine and an inverse association with DFI was found in Greenland, but not significant in the adjusted results (Supplementary Table 3). A sub-analysis with adjustment for all of the above mentioned covariates except for country showed results largely similar to the unadjusted results, indicating that adjustment for country was the main contributor to the differences between unadjusted and adjusted results (Supplementary Table 4). 4. Discussion The present study indicates no major effects of BDE-47 or BDE153 exposure on male reproductive function in the environmental exposure range found in the European and Arctic study populations. The exposure levels in the present study varied considerably across countries, with the highest serum level for the measured PBDEs in Greenlandic men. PBDEs have bioaccumulative potential and are found in high concentrations in marine food [6], and the main source of exposure in Greenland is probably through marine diet, whereas other sources of PBDE exposure such as house dust and indoor air may contribute to a larger extent to the levels found in Poland and Ukraine. A marine diet rich in omega-3 fatty acids has been indicated to be beneficial for semen quality [37] and it can be speculated that the predominantly marine diet in Greenland may counteract possible adverse effects on semen quality. A separate publication describes in detail possible determinants of PBDE, indicating poor correlation to most covariates, with country and area of living within Greenland being the strongest determinants of PBDE concentration in serum samples, and inconsistent associations with age in the different countries [38]. Similar exposure levels as identified in the present study have been found in other European studies [39]. However, it should be noted that the concentrations of BDE-47 and BDE-153 found in European countries and Greenland are considerably lower than concentrations in the United States [40]. Higher flammability standards in the United States are probably a main reason for the differences. At the time of sample collection for the present study, the production and use of PBDEs was not restricted, but during the past decade several regulatory initiatives and the addition of some PBDEs to the Stockholm Convention on Persistent Organic Pollutants have limited production and release of PBDEs into the environment. DNA damage has to our knowledge not been studied in previous epidemiological studies in relation to PBDE exposure, but in vitro studies on human neuroblastoma, and hepatoma cells indicate that BDE-47 may induce DNA damage as measured by the Comet assay [41,42], suggesting that similar damage to sperm cells may also occur.
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Table 1 Characteristics of the study populations represented as median and 5th; 95th percentiles for continuous outcomes and percentage of the study population with the characteristic for dichotomous outcomes. Characteristics
Greenland (n = 99)
Age (years) Abstinence time (days) Body mass index (kg/m2 ) BDE-47 ng/g lipid BDE-153 ng/g lipid Reported spillage (%) Previous urogenital infections (%) Smoking (%) Semen samples analyzed < 60 minutes after collection (%) Blood sampled before 12:00 (%)
32.1 (21.2, 43.6) 3.0 (1.0, 9.0) 25.5 (20.5, 31.8) 2.0 (0.6, 6.9) 2.7 (1.3, 7.8) 10 85 67 100 13
Poland (n = 100)
Ukraine (n = 100)
29.6 (25.3, 36.9) 3.0 (1.0, 30.0) 25.0 (20.2, 30.5) 0.6 (0.2, 5.3) 0.5 (0.3, 1.6) 6 5 28 63 95
26.1 (20.7, 38.2) 3.5 (1.8, 7.0) 23.6 (19.6, 29.4) 0.2 (0.0, 1.1) 0.3 (0.1, 0.8) 16 1 69 90 86
Table 2 Markers of male reproductive function in relation to BDE-47 exposure. Outcome
N
Crude ˇ (95% CI)
Adjusteda ˇ (95% CI)
Sperm conc. (106 /mL) Volume (mL)b Total count (×106 )b Motile sperm (%)c Normal cells (%) LH (IU/L) FSH (IU/L) Inhibin B (ng/L) SHBG (mmol/L) Testosterone Estradiol FAI DFI (%) HDS (%) TUNEL (%) FAS (%) Bcl-xL (%)
298 267 266 249 295 235 234 235 234 235 234 234 279 279 236 223 152
−0.01 (−0.06, 0.04) 0.01 (−0.02, 0.04) −0.02 (−0.08, 0.04) −0.01 (−0.03, 0.02) −0.01 (−0.05, 0.02) −0.01 (−0.03, 0.01) 0.01 (−0.02, 0.04) −0.02 (−0.04, −0.00)* −0.01 (−0.03, 0.02) −0.04 (−0.06, −0.02)* −0.03 (−0.05, −0.01)* −0.03 (−0.05, −0.01)* −0.05 (−0.09, −0.02)* −0.03 (−0.06, 0.01) −0.06 (−0.13, 0.01) 0.00 (−0.07, 0.08) 0.00 (−0.16, 0.15)
0.01 (−0.05, 0.06) 0.02 (−0.01, 0.06) 0.01 (−0.06, 0.08) 0.00 (−0.03, 0.04) −0.02 (−0.06, 0.02) −0.01 (−0.04, 0.01) 0.01 (−0.03, 0.04) −0.01 (−0.04, 0.01) −0.01 (−0.04, 0.01) −0.02 (−0.04, 0.00) −0.01 (−0.03, 0.01) −0.01 (−0.03, 0.02) −0.02 (−0.06, 0.02) −0.05 (−0.09, −0.01)* 0.04 (−0.04, 0.11) 0.00 (−0.09, 0.09) 0.15 (−0.03, 0.34)
*
Indicates significant association between BDE-47 and the outcome (p < 0.05). All results were adjusted for age, smoking, BMI and country. Sperm quality measures and apoptotic markers were additionally adjusted for abstinence time and urogenital infections. Reproductive hormones were additionally adjusted for blood sample before noon (12:00) or after. b Restricted to samples from men reporting no spillage. c Restricted to samples analyzed within one hour after semen collection. a
Table 3 Markers of male reproductive function in relation to BDE-153 exposure. Outcome
N
Crude ˇ (95% CI)
Adjusteda ˇ (95% CI)
Sperm conc. (106 /mL) Volume (mL)b Total count (×106 )b Motile sperm (%)c Normal cells (%) LH FSH Inhibin B SHBG Testosterone Estradiol FAI DFI HDS TUNEL FAS Bcl-xL
298 267 266 249 295 235 234 235 234 235 234 234 279 279 236 223 152
−0.04 (−0.13, 0.05) −0.00 (−0.06, 0.05) −0.07 (−0.19, 0.05) −0.03 (−0.08, 0.02) 0.01 (−0.06, 0.08) 0.02 (−0.03, 0.07) 0.08 (0.02, 0.14)* −0.03 (−0.07, 0.02) 0.06 (0.01, 0.10)* −0.05 (−0.09, −0.00)* −0.05 (−0.09, −0.02)* −0.10 (−0.14, −0.05)* −0.12 (−0.19, −0.06)* 0.04 (−0.02, 0.10) −0.37 (−0.48, −0.25)* −0.13 (−0.26, 0.01) −0.12 (−0.35, 0.10)
0.06 (−0.13, 0.24) −0.01 (−0.12, 0.11) 0.01 (−0.22, 0.24) 0.05 (−0.06, 0.16) 0.07 (−0.07, 0.20) 0.08 (−0.02, 0.18) 0.12 (−0.01, 0.25) −0.02 (−0.11, 0.07) 0.10 (−0.09, 0.30) 0.04 (−0.04, 0.12) 0.01 (−0.06, 0.09) −0.01 (−0.09, 0.07) −0.04 (−0.17, 0.09) −0.05 (−0.17, 0.07) −0.03 (−0.26, 0.21) −0.08 (−0.37, 0.21) −0.11 (−0.56, 0.34)
*
Indicates significant association between BDE-153 and the outcome (p < 0.05). All results were adjusted for age, smoking, BMI and country. Sperm quality measures and apoptopic markers were additionally adjusted for abstinence time and urogenital infections. Reproductive hormones were additionally adjusted for blood sample before noon (12:00) or after. b Restricted to samples from men reporting no spillage. c Restricted to samples analyzed within one hour after semen collection. a
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We did observe an inverse association between BDE-47 and HDS (a SCSA-derived parameter descriptive of sperm chromatin stability), but only marginally significant. We found no effects across countries on the more commonly used markers of sperm DNA damage (SCSA DFI and TUNEL) in the confounder adjusted results, indicating no strong effects of PBDEs on DNA damage. Furthermore, due to multiple testing it cannot be excluded that the association between BDE-47 and HDS and the effects only observed in single countries are chance findings. The main results of the present study do not corroborate previous findings of adverse effects of PBDEs on male sperm concentration, motility or reproductive hormones [17–20]. The previous studies on PBDE and male reproductive function were smaller (from 10 to 62 participants) and are therefore associated with larger uncertainties on the risk estimates. However, presumably higher exposure levels in the American studies [19,20], where PBDE exposure was measured in house dust, and in the Canadian study may explain some of the observed differences. However, in the Japanese study [17], similar exposure levels to PBDEs as in the present study were observed. A crucial difference between the studies is that the participants in the American and Canadian studies were recruited from infertility clinics, whereas the men in the present study are fertile men. Infertility patients may be more sensitive to exposure to PBDEs, but our populations may be a more representative sample of the general population, although infertile men are underrepresented. Although the cross-sectional nature of the present study limits the interpretation of the results, the long half-life of PBDEs [3] makes a single measurement of PBDE levels in serum collected close to the time of semen sampling a good estimate of the exposure for at least the duration of spermatogenesis in the testis (approximately 64 days) [43]. Although no effect of adult exposure to PBDEs on male reproductive function was observed, it can not be excluded that exposure in more sensitive time periods during organ development in the fetal period may have adverse effects on male reproductive function as indicated in recent studies with other environmental chemicals [44,45]. We adjusted the results in the present study for a number of potential confounders. A sub-analysis indicated that the main changes in estimates were observed upon including country in the model. We did observe several associations between PBDE exposure and reproductive hormones and sperm DNA structure in the models not adjusted for country. However, these associations were not observed at the individual country level, and although the individual countries may represent less contrast of exposure, we assume the most likely explanation is confounding by unmeasured factors or residual confounding by factors that differ between countries. Since the present study included data from different countries with greatly varying and only partly overlapping exposure levels, and also differences in some of the measured outcomes, adjustment for country is needed to limit the risk of confounding by unmeasured confounding factors associated with country. However, by adjusting for country we might overadjust our models, since we partly adjust for exposure, by including country in the model. Therefore, the model with the lowest risk of false positive associations would be the model adjusted for country, but in this model it cannot be excluded that we might overlook some of the associations that are only found by comparing the extreme ends of the exposure range. In conclusion our study indicated that PBDE exposure was not associated with a deterioration of semen quality and after adjustment for confounders no associations with reproductive hormones were found. Overall the results indicate no effect of environmental PBDE exposure on fertile adult male reproductive function in European and Arctic populations.
Acknowledgements The data on male fertility was collected as part of the INUENDO study (www.inuendo.dk) funded by the European Commission’s 5th Framework Program (QLK4-CT-2001-00202). Determination of PBDEs in serum and data analyses were part of the CLEAR study (www.inuendo.dk/clear) supported by the European Commission’s 7th Framework Program (FP7-ENV-2008-1-226217). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.reprotox. 2013.10.002. References [1] ATSDR. Toxicological profile for polybrominated biphenyls and polybrominated diphenyl ethers. U.S. Department of Health and Human Services, Agency for Toxic Substances and Disease Registry; 2004. [2] EFSA. EFSA panel on contaminants in the food chain (CONTAM) 2011 scientific opinion on polybrominated diphenyl ethers (PBDEs) in food. EFSA J 2011;9:2166. [3] Trudel D, Scheringer M, von Goetz N, Hungerbuhler K. Total consumer exposure to polybrominated diphenyl ethers in North America and Europe. Environ Sci Technol 2011;45:2391–7. [4] de Wit CA, Alaee M, Muir DC. Levels and trends of brominated flame retardants in the Arctic. Chemosphere 2006;64:209–33. [5] Frederiksen M, Thomsen C, Froshaug M, Vorkamp K, Thomsen M, Becher G, et al. Polybrominated diphenyl ethers in paired samples of maternal and umbilical cord blood plasma and associations with house dust in a Danish cohort. Int J Hyg Environ Health 2010;213:233–42. [6] Frederiksen M, Vorkamp K, Thomsen M, Knudsen LE. Human internal and external exposure to PBDEs – a review of levels and sources. Int J Hyg Environ Health 2009;212:109–34. [7] Leijs MM, ten Tusscher GW, Olie K, van Teunenbroek T, van Aalderen WM, de Voogt P, et al. Thyroid hormone metabolism and environmental chemical exposure. Environ Health 2012;11(Suppl 1):S10. [8] Stoker TE, Cooper RL, Lambright CS, Wilson VS, Furr J, Gray LE. In vivo and in vitro anti-androgenic effects of DE-71, a commercial polybrominated diphenyl ether (PBDE) mixture. Toxicol Appl Pharmacol 2005;207:78–88. [9] Mercado-Feliciano M, Bigsby RM. The polybrominated diphenyl ether mixture DE-71 is mildly estrogenic. Environ Health Perspect 2008;116:605–11. [10] Hamers T, Kamstra JH, Sonneveld E, Murk AJ, Kester MH, Andersson PL, et al. In vitro profiling of the endocrine-disrupting potency of brominated flame retardants. Toxicol Sci 2006;92:157–73. [11] van der Ven LT, van de KT, Verhoef A, Leonards PE, Slob W, Canton RF, et al. A 28-day oral dose toxicity study enhanced to detect endocrine effects of a purified technical pentabromodiphenyl ether (pentaBDE) mixture in Wistar rats. Toxicology 2008;245:109–22. [12] Muirhead EK, Skillman AD, Hook SE, Schultz IR. Oral exposure of PBDE-47 in fish: toxicokinetics and reproductive effects in Japanese Medaka (Oryzias latipes) and fathead minnows (Pimephales promelas). Environ Sci Technol 2006;40:523–8. [13] Marteinson SC, Kimmins S, Bird DM, Shutt JL, Letcher RJ, Ritchie IJ, et al. Embryonic exposure to the polybrominated diphenyl ether mixture, DE-71, affects testes and circulating testosterone concentrations in adult American kestrels (Falco sparverius). Toxicol Sci 2011;121:168–76. [14] Norris JM, Kociba RJ, Schwetz BA, Rose JQ, Humiston CG, Jewett GL, et al. Toxicology of octabromobiphenyl and decabromodiphenyl oxide. Environ Health Perspect 1975;11:153–61. [15] Tseng LH, Lee CW, Pan MH, Tsai SS, Li MH, Chen JR, et al. Postnatal exposure of the male mouse to 2,2 ,3,3 ,4,4 ,5,5 ,6,6 -decabrominated diphenyl ether: decreased epididymal sperm functions without alterations in DNA content and histology in testis. Toxicology 2006;224:33–43. [16] Tseng LH, Hsu PC, Lee CW, Tsai SS, Pan MH, Li MH. Developmental exposure to decabrominated diphenyl ether (BDE-209): effects on sperm oxidative stress and chromatin DNA damage in mouse offspring. Environ Toxicol 2013;28:380–9. [17] Akutsu K, Takatori S, Nozawa S, Yoshiike M, Nakazawa H, Hayakawa K, et al. Polybrominated diphenyl ethers in human serum and sperm quality. Bull Environ Contam Toxicol 2008;80:345–50. [18] Abdelouahab N, Ainmelk Y, Takser L. Polybrominated diphenyl ethers and sperm quality. Reprod Toxicol 2011;31:546–50. [19] Johnson PI, Stapleton HM, Mukherjee B, Hauser R, Meeker JD. Associations between brominated flame retardants in house dust and hormone levels in men. Sci Total Environ 2013;445-446:177–84. [20] Meeker JD, Johnson PI, Camann D, Hauser R. Polybrominated diphenyl ether (PBDE) concentrations in house dust are related to hormone levels in men. Sci Total Environ 2009;407:3425–9.
G. Toft et al. / Reproductive Toxicology 43 (2014) 1–7 [21] Toft G, Axmon A, Giwercman A, Thulstrup A, Rignell-Hydbom A, Pedersen HS, et al. Fertility in four regions spanning large contrasts in serum levels of widespread persistent organochlorines: a cross-sectional study. Environ Health 2005;4:26. [22] WHO. WHO laboratory manual for the examination of human semen and sperm-cervical mucus interaction. Cambridge, UK/New York: Cambridge University Press; 1999. [23] Toft G, Rignell-Hydbom A, Tyrkiel E, Shvets M, Giwercman A. Quality control workshops in standardization of sperm concentration and motility assessment in multicentre studies. Int J Androl 2005;28:144–9. [24] Evenson DP, Larson KL, Jost LK. Sperm chromatin structure assay: its clinical use for detecting sperm DNA fragmentation in male infertility and comparisons with other techniques. J Androl 2002;23:25–43. [25] Spano M, Toft G, Hagmar L, Eleuteri P, Rescia M, Rignell-Hydbom A, et al. Exposure to PCB and p,p -DDE in European and Inuit populations: impact on human sperm chromatin integrity. Hum Reprod 2005;20:3488–99. [26] Stronati A, Manicardi GC, Cecati M, Bordicchia M, Ferrante L, Spano M, et al. Relationships between sperm DNA fragmentation, sperm apoptotic markers and serum levels of CB-153 and p,p -DDE in European and Inuit populations. Reproduction 2006;132:949–58. [27] Groome NP, Illingworth PJ, O’Brien M, Pai R, Rodger FE, Mather JP, et al. Measurement of dimeric inhibin B throughout the human menstrual cycle. J Clin Endocrinol Metab 1996;81:1401–5. [28] Rignell-Hydbom A, Rylander L, Giwercman A, Jonsson BA, Lindh C, Eleuteri P, et al. Exposure to PCBs and p,p -DDE and human sperm chromatin integrity. Environ Health Perspect 2005;113:175–9. [29] Thomsen C, Liane VH, Becher G. Automated solid-phase extraction for the determination of polybrominated diphenyl ethers and polychlorinated biphenyls in serum – application on archived Norwegian samples from 1977 to 2003. J Chromatogr B: Analyt Technol Biomed Life Sci 2007;846:252–63. [30] Jönsson BAG, Rylander L, Lindh C, Rignell-Hydbom A, Giwercman A, Toft G, et al. Inter-population variations in concentrations, determinants of and correlations between 2,2 ,4,4 ,5,5 -hexachlorobiphenyl (CB-153) and 1,1-dichloro-2,2-bis (p-chlorophenyl)-ethylene (p,p -DDE): a cross-sectional study of 3161 men and women from Inuit and European populations. Environ Health 2005;4: 27. [31] Rylander L, Nilsson-Ehle P, Hagmar L. A simplified precise method for adjusting serum levels of persistent organohalogen pollutants to total serum lipids. Chemosphere 2006;62:333–6. [32] Lubin JH, Colt JS, Camann D, Davis S, Cerhan JR, Severson RK, et al. Epidemiologic evaluation of measurement data in the presence of detection limits. Environ Health Perspect 2004;112:1691–6.
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[33] Stewart TM, Liu DY, Garrett C, Jorgensen N, Brown EH, Baker HW. Associations between andrological measures, hormones and semen quality in fertile Australian men: inverse relationship between obesity and sperm output. Hum Reprod 2009;24:1561–8. [34] Molina RI, Martini AC, Tissera A, Olmedo J, Senestrari D, de Cuneo MF, et al. Semen quality and aging: analysis of 9.168 samples in Cordoba, Argentina. Arch Esp Urol 2010;63:214–22. [35] Ramlau-Hansen CH, Thulstrup AM, Aggerholm AS, Jensen MS, Toft G, Bonde JP. Is smoking a risk factor for decreased semen quality? A cross-sectional analysis. Hum Reprod 2007;22:188–96. [36] Ochsendorf FR. Sexually transmitted infections: impact on male fertility. Andrologia 2008;40:72–5. [37] Attaman JA, Toth TL, Furtado J, Campos H, Hauser R, Chavarro JE. Dietary fat and semen quality among men attending a fertility clinic. Hum Reprod 2012;27:1466–74. [38] Lenters V, Thomsen C, Smit LA, Jönsson BA, Pedersen HS, Ludwicki JK, et al. Serum concentrations of polybrominated diphenyl ethers (PBDEs) and a polybrominated biphenyl (PBB) in men from Greenland, Poland and Ukraine. Environ Int 2013;61C:8–16. [39] Leijs MM, van Teunenbroek T, Olie K, Koppe JG, ten Tusscher GW, van Aalderen WM, et al. Assessment of current serum levels of PCDD/Fs, dl-PCBs and PBDEs in a Dutch cohort with known perinatal PCDD/F exposure. Chemosphere 2008;73:176–81. [40] Sjodin A, Wong LY, Jones RS, Park A, Zhang Y, Hodge C, et al. Serum concentrations of polybrominated diphenyl ethers (PBDEs) and polybrominated biphenyl (PBB) in the United States population: 2003–2004. Environ Sci Technol 2008;42:1377–84. [41] Pellacani C, Buschini A, Galati S, Mussi F, Franzoni S, Costa LG. Evaluation of DNA damage induced by 2 polybrominated diphenyl ether flame retardants (BDE-47 and BDE-209) in SK-N-MC cells. Int J Toxicol 2012;31:372–9. [42] An J, Li S, Zhong Y, Wang Y, Zhen K, Zhang X, et al. The cytotoxic effects of synthetic 6-hydroxylated and 6-methoxylated polybrominated diphenyl ether 47 (BDE47). Environ Toxicol 2011;26:591–9. [43] Heller CG, Clermont Y. Spermatogenesis in man: an estimate of its duration. Science 1963;140:184–6. [44] Vested A, Ramlau-Hansen CH, Olsen SF, Bonde JP, Kristensen SL, Halldorsson TI, et al. Associations of in utero exposure to perfluorinated alkyl acids with human semen quality and reproductive hormones in adult men. Environ Health Perspect 2013;121:453–8. [45] Mocarelli P, Gerthoux PM, Needham LL, Patterson Jr DG, Limonta G, Falbo R, et al. Perinatal exposure to low doses of dioxin can permanently impair human semen quality. Environ Health Perspect 2011;119:713–8.
Contents lists available at ScienceDirect
Reproductive Toxicology journal homepage: www.elsevier.com/locate/reprotox
Exposure to polybrominated diphenyl ethers and male reproductive function in Greenland, Poland and Ukraine Gunnar Toft a,∗ , Virissa Lenters b , Roel Vermeulen b , Dick Heederik b , Cathrine Thomsen c , Georg Becher c,d , Aleksander Giwercman e , Davide Bizzaro f , Gian Carlo Manicardi g , k ´ Marcello Spanò h , Lars Rylander i , Henning S. Pedersen j , Paweł Strucinski , l m Valentyna Zviezdai , Jens Peter Bonde a
Danish Ramazzini Center, Department of Occupational Medicine, Aarhus University Hospital, Aarhus, Denmark Institute for Risk Assessment Sciences, Utrecht University, Utrecht, The Netherlands c Division of Environmental Medicine, Norwegian Institute of Public Health, Oslo, Norway d Department of Chemistry, University of Oslo, Oslo, Norway e Reproductive Medicine Centre, Skåne University Hospital, Lund University, Malmö, Sweden f Department of Life and Environmental Sciences, Polytechnic University of Marche, Ancona, Italy g Department of Life Science, Università di Modena e Reggio Emilia, Reggio Emilia, Italy h Laboratory of Toxicology, Unit of Radiation Biology and Human Health, ENEA Casaccia Research Center, Rome, Italy i Division of Occupational and Environmental Medicine, Lund University, Lund University Hospital, Lund, Sweden j Centre for Arctic Environmental Medicine, Nuuk, Greenland k Department of Toxicology and Risk Assessment, National Institute of Public Health–National Institute of Hygiene, Warsaw, Poland l Department of Social Medicine and Organization of Public Health, Kharkiv National Medical University, Kharkiv, Ukraine m Department of Occupational and Environmental Medicine, Bispebjerg University Hospital, Copenhagen, Denmark b
a r t i c l e
i n f o
Article history: Received 22 May 2013 Received in revised form 19 September 2013 Accepted 8 October 2013 Available online 26 October 2013 Keywords: Semen quality Reproductive hormones Inhibin B DNA damage Apoptosis PBDE BDE-47 BDE-153
a b s t r a c t Animal and a few human studies suggest that polybrominated diphenyl ethers (PBDEs) may affect male reproductive function. The aim of the present study was to evaluate if male reproductive function was associated with serum level of PBDEs. We evaluated, in a cross-sectional study, the effects of environmental exposure to BDE-47 and BDE-153 on reproductive hormones and semen quality, including markers of DNA damage and apoptosis, in 299 spouses of pregnant women from Greenland, Poland and Ukraine. Adjusted linear regression models indicated no strong associations between BDE-47 or BDE-153 exposure and markers of male semen quality or reproductive hormones. In the largest study to date we demonstrate that BDE-47 and BDE-153 exposure was not associated with altered semen characteristics or reproductive hormones, indicating that male reproductive function is not affected by the exposure level of these compounds in fertile European or Arctic populations. © 2013 Elsevier Inc. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs) have been used since the 1970s as additive flame retardants in a variety of consumer products including polyurethane foam used in mattresses and upholstered furniture and thermoplastics used in electronic
∗ Corresponding author at: Danish Ramazzini Center, Department of Occupational Medicine, Aarhus University Hospital, Norrebrogade 44, Building 2C, 8000 Aarhus C, Denmark. Tel.: +45 78 46 4251; fax: +45 7846 4260. E-mail address: [email protected] (G. Toft). 0890-6238/$ – see front matter © 2013 Elsevier Inc. All rights reserved. http://dx.doi.org/10.1016/j.reprotox.2013.10.002
equipment [1]. As they are mixed into polymers, and not chemically bound to the plastics or textiles, they are able to separate or leach from the products into the environment [2]. PBDEs are lipophilic compounds with a biological half-life up to 7 years [3]. Important sources of exposure include fatty fish and mammals from higher trophic levels, and therefore health risks for Arctic populations are a special concern [4]. House dust has also been indicated as an important contributor to human PBDE exposure, especially for toddlers and small children [5,6]. Of the 209 PBDE congeners, BDE47 is usually the compound detected in the highest concentration in human biological material, followed by BDE-153 [5]. In addition to effects on the thyroid hormones [7], several PBDEs have shown the
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potential to interfere with reproductive hormone signaling through anti-androgenic, estrogenic or anti-estrogenic activity in in vitro models [8–10]. In vivo subacute (28 day) toxicity tests on adult male rats exposed to a commercial pentaBDE mixture in a dose range from 0.27 to 200 mg/kg body weight/day have shown dosedependent decreased weight of the epididymis, seminal vesicles and prostate, and an increase in sperm head deformities with lower confidence bound of the estimated benchmark dose in the range of 10–50 mg/kg body weight/day [11]. Exposure of the fish Fathead Minnows (Pimephales promelas) to 14 g BDE-47/day for 25 days indicated specific effects on male reproductive function, including a reduction in sperm concentration [12]. Furthermore, exposure in ovo of American kestrels (Falco sparverius) to the commercial pentaBDE mixture, DE-71, measured at 1 g/g wet weight in the eggs, reduced testosterone levels and altered testis structure in the male offspring [13]. Some earlier toxicological evaluations of PBDEs did not indicate adverse effects on reproductive function in rodents after exposure in adulthood [14], but more recent experiments in mice have reported that, whereas neonatal exposure to 500 or 1500 mg/kg BDE-209 per day could induce weak effects on a few aspects of sperm function [15], in utero exposure to BDE209 induced testicular changes together with a significant increase of sperm head abnormalities and sperm chromatin damage in the male progeny [16], and some of these effects were observed at a much lower concentration (10 mg/kg/day BDE-209). In humans, only a few studies addressed the issue of environmental PBDE exposure in relation to adult male reproductive function. A Japanese study of 10 young men indicated adverse associations between BDE-153 and sperm concentration and testis size [17], and a Canadian study of 52 men recruited in an infertility clinic indicated adverse associations of BDE-47, BDE-100 and the sum of measured PBDE congeners (BDE-47, BDE-99, BDE-100 and BDE-153) on sperm motility [18], but no relation with other semen parameters. Finally, an American study of 62 men indicated positive associations between house dust penta-BDEs and serum level of estradiol and sex hormone binding globulin (SHBG), and inverse associations with follicle stimulating hormone (FSH). In the same study, house dust octa-BDEs were positively associated with luteinizing hormone (LH) and testosterone and finally deca-BDEs were inversely associated with testosterone [19]. This study was expanding on a previous report from the same group indicating an inverse associations between the measured PBDE congeners (47, 99 and 100) in dust samples and free testosterone, LH and FSH [20]. Thus, although epidemiological studies are only partly supported by experimental studies, animal and human studies suggest that PBDEs may affect human reproductive health. The associations have so far only been evaluated in small studies with divergent endpoints, and confirmation is needed in larger studies before more definite conclusions on effects on male reproductive health can be made. The aim of the present study was to evaluate if PBDE exposure is associated with male reproductive function in three populations of fertile men with considerable variation in PBDE exposure.
2. Materials and methods 2.1. Study populations The present study is part of a European study on fertility, the CLEAR study (www.inuendo.dk/clear), using a uniform protocol for data collection in Greenland, Kharkiv in Ukraine, and Warsaw in Poland [21]. Six hundred and two partners of pregnant women provided a semen sample and a blood sample, and filled in a questionnaire on lifestyle, occupation and medical history. All men were 18 years
or older at the time of enrolment. In each of the three countries 100 blood samples were selected randomly among the samples with sufficient volume to allow analysis (>4 mL). One sample was lost due to problems during sample preparation before chemical analyses, and therefore the final sample size used to evaluate the effect of lipid adjusted PBDEs on male reproductive outcomes was 299. The 299 included men were not significantly different from the 303 non-included men regarding measures of semen quality or reproductive hormones. The participation rates in the study varied from 29% in Warsaw and 33% in Kharkiv to 79% in Greenland. Study populations and data collection procedures have been described in detail elsewhere [21]. 2.2. Collection of semen samples and basic semen analysis All semen samples were collected by masturbation. The subject was asked to abstain from sexual activities for at least two days before collecting the sample, and to report the actual abstinence time. The sample was kept close to the body to maintain a temperature close to 37 ◦ C if transport to the local hospital after collection was necessary. Analysis of semen samples was initiated within one hour after ejaculation for 83% of the samples. The samples were analyzed for concentration, motility and morphology according to a manual for the project based on the 1999 World Health Organization (WHO) manual for basic semen analysis [22], which was the most recent version of the manual at the time of sample collection. Briefly, sperm concentration was determined in duplicate using an Improved Neubauer Hemacytometer. Sperm motility was determined by counting the proportion of (a) rapid progressive spermatozoa; (b) slow progressive spermatozoa versus (c) nonprogressive motile spermatozoa; and (d) immotile spermatozoa among 100 spermatozoa within each of two fresh drops of semen. Semen samples from Warsaw and Kharkiv were analyzed for concentration and motility at one central hospital in each region, whereas the samples from Greenland were analyzed at the local hospitals or nursing stations spread across the country. One person performed all the semen analyses in each of the three countries. These three persons were previously trained in a quality control program set up specifically for this study [23]. The inter-observer variation as consider concentration was found to be 8% and for motility 11%. Spermatozoa from all populations were stained and analyzed for morphology centrally at Skåne University Hospital in Malmö, Sweden. Abnormalities were classified as head defects, midpiece defects, tail defects, cytoplasma drop and immature spermatozoa. The morphology was evaluated for at least 200 sperms in each sample by two technicians taking part in the NAFA-ESHRE (Nordic Association for Andrology – European Society of Human Reproduction and Embryology) external quality control program. 2.3. Sperm chromatin structure assay (SCSA) DNA damage was measured by SCSA following the standardized procedure described by Evenson et al. [24]. Briefly, all coded frozen semen samples from the three study regions were shipped on dry ice to the flow cytometry facility of the ENEA Laboratory of Toxicology (Rome, Italy) for SCSA analysis; a flow cytometric (FCM) technique which identifies the spermatozoa with abnormal chromatin packaging, envisioned by increased susceptibility to acid-induced DNA denaturation in situ followed by acridine orange fluorescence staining [24,25]. Adhering strictly to the SCSA protocol as described by Evenson et al. [24], the DNA fragmentation index (%DFI) – representing the percentage of sperm with detectable DNA breaks – together with the percentage of sperm with high levels of high DNA stainability (%HDS) – representing the fraction of sperm with anomalies in the histone-to-protamine transition – were
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calculated by a dedicated software (SCSASoft; SCSA Diagnostics, Volga, SD, USA). FCM measurements were stopped when a total of 10,000 spermatozoa had been accumulated for each sample. For the flow cytometer set-up and calibration, a reference semen sample retrieved from our laboratory repository was used as described in [25]. Inter-day SCSA variability for %DFI, based on 216 flow sessions showed a coefficients of variation (CV) of 6.0%, whereas the CV for the %HDS was 4.8%. In addition, 358 randomly chosen samples (50.6% of the total) were measured twice in independent FCM sessions. Results from the two measurements were highly correlated (DFI, r = 0.96). External quality control indicated a median inter-sample variability for %DFI, expressed as CV, of 1.5% [25]. 2.4. Determination of DNA fragmentation by terminal deoxynucleotidyl transferase dUTP nick end labeling (TUNEL) assay
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0.2 IU/L, 0.2 IU/L, and 8.0 pmol/L, respectively. The total assay coefficients of variation were 2.9%, 2.6%, and 8.1%, respectively. Serum testosterone levels were measured by means of a competitive immunoassay (Access; Beckman Coulter Inc., Fullerton, CA, USA) with an LOD of 0.35 nmol/L and a total assay CV of 2.8% at 2.9 nmol/L and 3.2% at 8.1 nmol/L. SHBG concentrations were measured using a fluoro-immunoassay (Immulite 2000; Diagnostic Products Corporation, Los Angeles, CA, USA). The LOD was 0.02 nmol/L. The total assay coefficients of variation were 5.5% and 4.6%, respectively. Inhibin B levels were assessed using a specific immunometric method, with a detection limit of 15 ng/L and intraassay and total assay coefficients of variation 40%), so the median blank contamination level of 1.58 pg/g serum was subtracted from all individual measurements. BFR concentrations were analyzed as lipid adjusted values. Serum concentrations of triglycerides and cholesterol were determined by enzymatic methods [30]. The total lipid concentration in serum (g/L) was calculated as
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total = 0.96 + 1.28 × (triglycerides + cholesterol), as described in [31]. 2.9. Statistical analysis We evaluated trends of exposure-outcome associations by general linear regression models across the three regions. Only exposures with measurements above the LOQ in more than 70% of the study population samples were included in statistical analyses [32], limiting the analyses to BDE-47 and BDE 153, detected above LOQ in 99% and 97% of samples, respectively. Values for samples below LOQ for these compounds were estimated and included in the statistical analysis. The values were estimated using iterative single imputation, in which the estimated means depended on the other measured PBDE exposures and study population to allow for population-specific residual variances [32]. Due to the skewed exposure distributions, we transformed PBDEs by taking the natural logarithm (ln). To improve normality and homogeneity of variance of the residuals in the tested associations, all outcome variables were ln-transformed. Model fit was evaluated by visual inspection of residuals, i.e. observed versus predicted values. Semen variables (sperm concentration, total sperm count, volume, motility, morphology, DFI, HDS, TUNEL, FAS and BclxL) were, according to an a priori decision, adjusted for potential confounding effects of known determinants of male reproductive function, including for all semen variables: abstinence time and age (both ln-transformed); current smoking (yes/no); ever urogenital infections (yes/no); body mass index (BMI – kg/m2 ) (25), and country [21,33–36]. Sperm concentration, motility and morphology were additionally adjusted for spillage (yes/no). Furthermore, analysis of sperm motility was restricted to samples analyzed within one hour of collection (n = 250), and volume and total sperm count were restricted to samples from individuals reporting no spillage (n = 267). Associations between PBDEs and reproductive hormones were adjusted for age, smoking, BMI, country and blood sampled before noon (12:00) or after. Effect modification by country was tested by interaction tests, by including a country × exposure term in the regression models including covariates as specified in the adjusted regression models for each exposure-outcome association tested. Exposure-outcome associations were additionally evaluated stratified by country in a supplementary analysis. In addition, supplementary analyses were run without country in the models. Statistical analyses were performed using the SAS statistical software version 9.1 (SAS Institute Inc., Cary, NC, USA). 3. Results 3.1. Characteristics of the study populations In total 299 men delivered a semen sample and had a blood sample analyzed for PBDEs. The median PBDE concentrations were up to 10 fold higher in serum samples from men from Greenland compared to Ukraine, with intermediate levels of the measured compounds in Poland. Also, other characteristics like age, percentage of persons reporting spillage, percentage reporting previous urogenital infections and percentage of the populations smoking varied considerably across countries (Table 1). 3.2. Associations between PBDE exposure and male reproductive function Markers of male reproductive function in relation to BDE-47 and BDE-153 are shown in Tables 2 and 3. In crude models, the measured PBDE congeners were inversely associated with estradiol, testosterone, FAI, and one marker of sperm DNA damage (DFI). In
addition BDE-47 was inversely associated with inhibin B, and BDE153 was positively associated with FSH, SHBG and DNA damage measured by the TUNEL assay. After adjustment for age, abstinence time, smoking, urogenital infections, BMI and country, none of the negative associations between BDE-47 or BDE-153 exposure and male reproductive health outcomes remained. However, we observed an inverse association between BDE-47 exposure and HDS after adjustment ˇ = −0.05 CI (−0.09; −0.01). No significant effect modification by country was indicated. Results stratified by country are shown in Supplementary Tables 2 and 3. Also in the stratified analyses, an inverse association between BDE-47 and HDS was found in all countries, although only statistically significant in Ukraine. Apart from this no consistent associations were found across countries. However, a few significant associations in single countries were observed, including for BDE-47: an inverse association with semen volume in Greenland but a positive association with semen volume in Poland, and a positive association with TUNEL in Poland and a positive association with Bcl-xL in Ukraine (Supplementary Table 2). For BDE-153, a positive association with FSH was observed in Ukraine and an inverse association with DFI was found in Greenland, but not significant in the adjusted results (Supplementary Table 3). A sub-analysis with adjustment for all of the above mentioned covariates except for country showed results largely similar to the unadjusted results, indicating that adjustment for country was the main contributor to the differences between unadjusted and adjusted results (Supplementary Table 4). 4. Discussion The present study indicates no major effects of BDE-47 or BDE153 exposure on male reproductive function in the environmental exposure range found in the European and Arctic study populations. The exposure levels in the present study varied considerably across countries, with the highest serum level for the measured PBDEs in Greenlandic men. PBDEs have bioaccumulative potential and are found in high concentrations in marine food [6], and the main source of exposure in Greenland is probably through marine diet, whereas other sources of PBDE exposure such as house dust and indoor air may contribute to a larger extent to the levels found in Poland and Ukraine. A marine diet rich in omega-3 fatty acids has been indicated to be beneficial for semen quality [37] and it can be speculated that the predominantly marine diet in Greenland may counteract possible adverse effects on semen quality. A separate publication describes in detail possible determinants of PBDE, indicating poor correlation to most covariates, with country and area of living within Greenland being the strongest determinants of PBDE concentration in serum samples, and inconsistent associations with age in the different countries [38]. Similar exposure levels as identified in the present study have been found in other European studies [39]. However, it should be noted that the concentrations of BDE-47 and BDE-153 found in European countries and Greenland are considerably lower than concentrations in the United States [40]. Higher flammability standards in the United States are probably a main reason for the differences. At the time of sample collection for the present study, the production and use of PBDEs was not restricted, but during the past decade several regulatory initiatives and the addition of some PBDEs to the Stockholm Convention on Persistent Organic Pollutants have limited production and release of PBDEs into the environment. DNA damage has to our knowledge not been studied in previous epidemiological studies in relation to PBDE exposure, but in vitro studies on human neuroblastoma, and hepatoma cells indicate that BDE-47 may induce DNA damage as measured by the Comet assay [41,42], suggesting that similar damage to sperm cells may also occur.
G. Toft et al. / Reproductive Toxicology 43 (2014) 1–7
5
Table 1 Characteristics of the study populations represented as median and 5th; 95th percentiles for continuous outcomes and percentage of the study population with the characteristic for dichotomous outcomes. Characteristics
Greenland (n = 99)
Age (years) Abstinence time (days) Body mass index (kg/m2 ) BDE-47 ng/g lipid BDE-153 ng/g lipid Reported spillage (%) Previous urogenital infections (%) Smoking (%) Semen samples analyzed < 60 minutes after collection (%) Blood sampled before 12:00 (%)
32.1 (21.2, 43.6) 3.0 (1.0, 9.0) 25.5 (20.5, 31.8) 2.0 (0.6, 6.9) 2.7 (1.3, 7.8) 10 85 67 100 13
Poland (n = 100)
Ukraine (n = 100)
29.6 (25.3, 36.9) 3.0 (1.0, 30.0) 25.0 (20.2, 30.5) 0.6 (0.2, 5.3) 0.5 (0.3, 1.6) 6 5 28 63 95
26.1 (20.7, 38.2) 3.5 (1.8, 7.0) 23.6 (19.6, 29.4) 0.2 (0.0, 1.1) 0.3 (0.1, 0.8) 16 1 69 90 86
Table 2 Markers of male reproductive function in relation to BDE-47 exposure. Outcome
N
Crude ˇ (95% CI)
Adjusteda ˇ (95% CI)
Sperm conc. (106 /mL) Volume (mL)b Total count (×106 )b Motile sperm (%)c Normal cells (%) LH (IU/L) FSH (IU/L) Inhibin B (ng/L) SHBG (mmol/L) Testosterone Estradiol FAI DFI (%) HDS (%) TUNEL (%) FAS (%) Bcl-xL (%)
298 267 266 249 295 235 234 235 234 235 234 234 279 279 236 223 152
−0.01 (−0.06, 0.04) 0.01 (−0.02, 0.04) −0.02 (−0.08, 0.04) −0.01 (−0.03, 0.02) −0.01 (−0.05, 0.02) −0.01 (−0.03, 0.01) 0.01 (−0.02, 0.04) −0.02 (−0.04, −0.00)* −0.01 (−0.03, 0.02) −0.04 (−0.06, −0.02)* −0.03 (−0.05, −0.01)* −0.03 (−0.05, −0.01)* −0.05 (−0.09, −0.02)* −0.03 (−0.06, 0.01) −0.06 (−0.13, 0.01) 0.00 (−0.07, 0.08) 0.00 (−0.16, 0.15)
0.01 (−0.05, 0.06) 0.02 (−0.01, 0.06) 0.01 (−0.06, 0.08) 0.00 (−0.03, 0.04) −0.02 (−0.06, 0.02) −0.01 (−0.04, 0.01) 0.01 (−0.03, 0.04) −0.01 (−0.04, 0.01) −0.01 (−0.04, 0.01) −0.02 (−0.04, 0.00) −0.01 (−0.03, 0.01) −0.01 (−0.03, 0.02) −0.02 (−0.06, 0.02) −0.05 (−0.09, −0.01)* 0.04 (−0.04, 0.11) 0.00 (−0.09, 0.09) 0.15 (−0.03, 0.34)
*
Indicates significant association between BDE-47 and the outcome (p < 0.05). All results were adjusted for age, smoking, BMI and country. Sperm quality measures and apoptotic markers were additionally adjusted for abstinence time and urogenital infections. Reproductive hormones were additionally adjusted for blood sample before noon (12:00) or after. b Restricted to samples from men reporting no spillage. c Restricted to samples analyzed within one hour after semen collection. a
Table 3 Markers of male reproductive function in relation to BDE-153 exposure. Outcome
N
Crude ˇ (95% CI)
Adjusteda ˇ (95% CI)
Sperm conc. (106 /mL) Volume (mL)b Total count (×106 )b Motile sperm (%)c Normal cells (%) LH FSH Inhibin B SHBG Testosterone Estradiol FAI DFI HDS TUNEL FAS Bcl-xL
298 267 266 249 295 235 234 235 234 235 234 234 279 279 236 223 152
−0.04 (−0.13, 0.05) −0.00 (−0.06, 0.05) −0.07 (−0.19, 0.05) −0.03 (−0.08, 0.02) 0.01 (−0.06, 0.08) 0.02 (−0.03, 0.07) 0.08 (0.02, 0.14)* −0.03 (−0.07, 0.02) 0.06 (0.01, 0.10)* −0.05 (−0.09, −0.00)* −0.05 (−0.09, −0.02)* −0.10 (−0.14, −0.05)* −0.12 (−0.19, −0.06)* 0.04 (−0.02, 0.10) −0.37 (−0.48, −0.25)* −0.13 (−0.26, 0.01) −0.12 (−0.35, 0.10)
0.06 (−0.13, 0.24) −0.01 (−0.12, 0.11) 0.01 (−0.22, 0.24) 0.05 (−0.06, 0.16) 0.07 (−0.07, 0.20) 0.08 (−0.02, 0.18) 0.12 (−0.01, 0.25) −0.02 (−0.11, 0.07) 0.10 (−0.09, 0.30) 0.04 (−0.04, 0.12) 0.01 (−0.06, 0.09) −0.01 (−0.09, 0.07) −0.04 (−0.17, 0.09) −0.05 (−0.17, 0.07) −0.03 (−0.26, 0.21) −0.08 (−0.37, 0.21) −0.11 (−0.56, 0.34)
*
Indicates significant association between BDE-153 and the outcome (p < 0.05). All results were adjusted for age, smoking, BMI and country. Sperm quality measures and apoptopic markers were additionally adjusted for abstinence time and urogenital infections. Reproductive hormones were additionally adjusted for blood sample before noon (12:00) or after. b Restricted to samples from men reporting no spillage. c Restricted to samples analyzed within one hour after semen collection. a
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G. Toft et al. / Reproductive Toxicology 43 (2014) 1–7
We did observe an inverse association between BDE-47 and HDS (a SCSA-derived parameter descriptive of sperm chromatin stability), but only marginally significant. We found no effects across countries on the more commonly used markers of sperm DNA damage (SCSA DFI and TUNEL) in the confounder adjusted results, indicating no strong effects of PBDEs on DNA damage. Furthermore, due to multiple testing it cannot be excluded that the association between BDE-47 and HDS and the effects only observed in single countries are chance findings. The main results of the present study do not corroborate previous findings of adverse effects of PBDEs on male sperm concentration, motility or reproductive hormones [17–20]. The previous studies on PBDE and male reproductive function were smaller (from 10 to 62 participants) and are therefore associated with larger uncertainties on the risk estimates. However, presumably higher exposure levels in the American studies [19,20], where PBDE exposure was measured in house dust, and in the Canadian study may explain some of the observed differences. However, in the Japanese study [17], similar exposure levels to PBDEs as in the present study were observed. A crucial difference between the studies is that the participants in the American and Canadian studies were recruited from infertility clinics, whereas the men in the present study are fertile men. Infertility patients may be more sensitive to exposure to PBDEs, but our populations may be a more representative sample of the general population, although infertile men are underrepresented. Although the cross-sectional nature of the present study limits the interpretation of the results, the long half-life of PBDEs [3] makes a single measurement of PBDE levels in serum collected close to the time of semen sampling a good estimate of the exposure for at least the duration of spermatogenesis in the testis (approximately 64 days) [43]. Although no effect of adult exposure to PBDEs on male reproductive function was observed, it can not be excluded that exposure in more sensitive time periods during organ development in the fetal period may have adverse effects on male reproductive function as indicated in recent studies with other environmental chemicals [44,45]. We adjusted the results in the present study for a number of potential confounders. A sub-analysis indicated that the main changes in estimates were observed upon including country in the model. We did observe several associations between PBDE exposure and reproductive hormones and sperm DNA structure in the models not adjusted for country. However, these associations were not observed at the individual country level, and although the individual countries may represent less contrast of exposure, we assume the most likely explanation is confounding by unmeasured factors or residual confounding by factors that differ between countries. Since the present study included data from different countries with greatly varying and only partly overlapping exposure levels, and also differences in some of the measured outcomes, adjustment for country is needed to limit the risk of confounding by unmeasured confounding factors associated with country. However, by adjusting for country we might overadjust our models, since we partly adjust for exposure, by including country in the model. Therefore, the model with the lowest risk of false positive associations would be the model adjusted for country, but in this model it cannot be excluded that we might overlook some of the associations that are only found by comparing the extreme ends of the exposure range. In conclusion our study indicated that PBDE exposure was not associated with a deterioration of semen quality and after adjustment for confounders no associations with reproductive hormones were found. Overall the results indicate no effect of environmental PBDE exposure on fertile adult male reproductive function in European and Arctic populations.
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